advances in oceanography and limnology advances in oceanography and limnology advances in oceanography and limnology advances in oceanography and limnology advances in oceanography and limnology advances in oceanography and limnology advances in oceanography and limnology advances in oceanography and limnology advances in oceanography and limnology advances in oceanography and limnology advances in oceanography and limnology advances in oceanography and limnology advances in oceanography and limnology advances in oceanography and limnology advances in oceanography and limnology advances in oceanography and limnology advances in oceanography and limnology advances in oceanography and limnology advances in oceanography and limnology advances in oceanography and limnology advances in oceanography and limnology advances in oceanography and limnology advances in oceanography and limnology advances in oceanography and limnology advances in oceanography and limnology advances in oceanography and limnology advances in oceanography and limnology 0.09 ky369673_1_limnodrilus_hoffmeisteriix_cnas51 ln810312_1_limnodrilus_hoffmeisterit20_262 ky369721_1_limnodrilus_hoffmeisterivi_cnl75 ky369719_1_limnodrilus_hoffmeisterivi_cnl59 ky369605_1_limnodrilus_hoffmeisterix_ce10835 ky369581_1_limnodrilus_hoffmeisterivii_ce8600 ky369627_1_limnodrilus_hoffmeisteriix_ce22811 ln999364_1_limnodrilus_hoffmeisteri668_t17 ln999368_1_limnodrilus_hoffmeisteri673_t21 ky369687_1_limnodrilus_hoffmeisterivi_cnh19 ky369543_1_limnodrilus_hoffmeisteriviii_ce4295 ky369481_1_limnodrilus_hoffmeisteriix_ce1843 mt450682_1_limnodrilus_hoffmeisteri_cw0017 ln810411_1_limnodrilus_hoffmeisterit18_264 9450_limnodrilus_hoffmeisteri ln999324_1_limnodrilus_hoffmeisteri617_t21 ln810404_1_limnodrilus_hoffmeisterit17_298 lr025081_1_limnodrilus_hoffmeisteri1145_t17 ky369626_1_limnodrilus_hoffmeisteriix_ce22809 ln999332_1_limnodrilus_hoffmeisteri625_t17 ky369679_1_limnodrilus_hoffmeisterivi_cnh3 ky369563_1_limnodrilus_hoffmeisterix_ce6577 ky369589_1_limnodrilus_hoffmeisterix_ce9004 ky369498_1_limnodrilus_hoffmeisteriviii_ce2721 ky369554_1_limnodrilus_hoffmeisteriiii_ce5203 ln999137_1_limnodrilus_hoffmeisteri431_t21 ky369585_1_limnodrilus_hoffmeisterii_ce8604 lt905369_1_limnodrilus_hoffmeisteri757_t18 ky369702_1_limnodrilus_hoffmeisteriix_cnj7 ky369609_1_limnodrilus_hoffmeisteriix_ce12492 ky369658_1_limnodrilus_hoffmeisterivii_ce22911 ln810397_1_limnodrilus_hoffmeisterit17_118 ky369714_1_limnodrilus_hoffmeisterivi_cnl9 ln999246_1_limnodrilus_hoffmeisteri539_t17 ln810387_1_limnodrilus_hoffmeisterit17_173 ln999226_1_limnodrilus_hoffmeisteri523_t17 ky369727_1_limnodrilus_hoffmeisterivi_cnwq13 ln999355_1_limnodrilus_hoffmeisteri654_t18 ln999219_1_limnodrilus_hoffmeisteri516_t17 9450_l005_limnodrilus_hoffmeisteri ky369590_1_limnodrilus_hoffmeisterivii_ce9014 ln810307_1_limnodrilus_hoffmeisterit21_265 ln810400_1_limnodrilus_hoffmeisterit17_271 ky369500_1_limnodrilus_hoffmeisteriviii_ce2739 ky369756_1_limnodrilus_hoffmeisterivi_cnyp32 ky369652_1_limnodrilus_hoffmeisteriii_ce22904 ln999073_1_limnodrilus_hoffmeisteri368_t20 ky369676_1_limnodrilus_hoffmeisterivi_cnf4 ky369528_1_limnodrilus_hoffmeisteriviii_ce3139 ky369587_1_limnodrilus_hoffmeisterii_ce8611 9450_pl05_limnodrilus_hoffmeisteri ln999255_1_limnodrilus_hoffmeisteri548_t17 ky369485_1_limnodrilus_hoffmeisterix_ce1991 ky369644_1_limnodrilus_hoffmeisterivii_ce22894 ky369608_1_limnodrilus_hoffmeisteriviii_ce10986 ky369572_1_limnodrilus_hoffmeisterivii_ce8564 ky369515_1_limnodrilus_hoffmeisteriix_ce3088 ky369700_1_limnodrilus_hoffmeisteriix_cnj4 ln810407_1_limnodrilus_hoffmeisterit17_114 ln810389_1_limnodrilus_hoffmeisterit17_167 ky369745_1_limnodrilus_hoffmeisterivi_cnyd32 ky369533_1_limnodrilus_hoffmeisterii_ce3242 mt450684_1_limnodrilus_hoffmeisteri_cw0022 lt899878_1_limnodrilus_hoffmeisteri859_t20 ky369499_1_limnodrilus_hoffmeisterivii_ce2722 ky369604_1_limnodrilus_hoffmeisteriv_ce10834 ky369724_1_limnodrilus_hoffmeisterivi_cnwq10 ky369736_1_limnodrilus_hoffmeisterivi_cnyd2 ky369648_1_limnodrilus_hoffmeisterivii_ce22899 ky369640_1_limnodrilus_hoffmeisteriii_ce22890 ky369637_1_limnodrilus_hoffmeisterivii_ce22886 ky369659_1_limnodrilus_hoffmeisteriii_ce22913 ln999165_1_limnodrilus_hoffmeisteri459_t17 lt903782_1_limnodrilus_hoffmeisteri999_t17 ky369560_1_limnodrilus_hoffmeisterix_ce5863 ky369670_1_limnodrilus_hoffmeisteriix_cnas14 ky369518_1_limnodrilus_hoffmeisteriix_ce3091 ky369519_1_limnodrilus_hoffmeisterii_ce3092 ky369752_1_limnodrilus_hoffmeisterivi_cnyp8 ky369755_1_limnodrilus_hoffmeisterivi_cnyp30 ky369665_1_limnodrilus_hoffmeisterix_cnas5 9450_a004_limnodrilus_hoffmeisteri ky369502_1_limnodrilus_hoffmeisteriviii_ce2741 ln999113_1_limnodrilus_hoffmeisteri401_t17 ky369722_1_limnodrilus_hoffmeisterivi_cnl80 ky369592_1_limnodrilus_hoffmeisterix_ce9773 ln999153_1_limnodrilus_hoffmeisteri444_t20 ln999341_1_limnodrilus_hoffmeisteri638_t21 lt899867_1_limnodrilus_hoffmeisteri842_t20 lt905376_1_limnodrilus_hoffmeisteri766_t18 ky369494_1_limnodrilus_hoffmeisterivii_ce2711 ky369738_1_limnodrilus_hoffmeisterivi_cnyd4 ky369643_1_limnodrilus_hoffmeisterivii_ce22893 ky369469_1_limnodrilus_hoffmeisterivii_ce1157 ln999124_1_limnodrilus_hoffmeisteri417_t21 ky369607_1_limnodrilus_hoffmeisteri_ce10860 ky369603_1_limnodrilus_hoffmeisterix_ce10783 ky369512_1_limnodrilus_hoffmeisterivii_ce2979 lt905368_1_limnodrilus_hoffmeisteri755_t17 ln810304_1_limnodrilus_hoffmeisterit21_127 ln810395_1_limnodrilus_hoffmeisterit17_305 ky369530_1_limnodrilus_hoffmeisterii_ce3239 ky369565_1_limnodrilus_hoffmeisterivii_ce6579 ky369559_1_limnodrilus_hoffmeisterivii_ce5862 9450_lh01_limnodrilus_hoffmeisteri ln999267_1_limnodrilus_hoffmeisteri560_t17 ky369677_1_limnodrilus_hoffmeisterivi_cnf5 ky369550_1_limnodrilus_hoffmeisterivii_ce5199 ln810391_1_limnodrilus_hoffmeisterit17_117 ln999099_1_limnodrilus_hoffmeisteri377_t17 lt899869_1_limnodrilus_hoffmeisteri850_t18 ln999188_1_limnodrilus_hoffmeisteri482_t17 mt460098_1_limnodrilus_profundicola_cw0152 ln810393_1_limnodrilus_hoffmeisterit17_306 ln999329_1_limnodrilus_hoffmeisteri622_t21 ky369647_1_limnodrilus_hoffmeisterivii_ce22898 ky369463_1_limnodrilus_hoffmeisterivii_ce604 ky369655_1_limnodrilus_hoffmeisteriii_ce22907 ln999204_1_limnodrilus_hoffmeisteri501_t17 ln999128_1_limnodrilus_hoffmeisteri421_t17 ky369531_1_limnodrilus_hoffmeisterii_ce3240 ln999175_1_limnodrilus_hoffmeisteri467_t18 lt905399_1_limnodrilus_hoffmeisteri836_t18 lt899883_1_limnodrilus_hoffmeisteri787_t18 ln999215_1_limnodrilus_hoffmeisteri512_t17 lt598634_1_limnodrilus_hoffmeisteri933 ky369505_1_limnodrilus_hoffmeisterivii_ce2759 ky369566_1_limnodrilus_hoffmeisterix_ce6581 ln810305_1_limnodrilus_hoffmeisterit21_334 ln999185_1_limnodrilus_hoffmeisteri478_t17 ln999307_1_limnodrilus_hoffmeisteri599_t21 ky369471_1_limnodrilus_hoffmeisterivii_ce1159 ky369616_1_limnodrilus_hoffmeisterix_ce14185 ky369595_1_limnodrilus_hoffmeisteriviii_ce10614 ln999150_1_limnodrilus_hoffmeisteri438_t21 ky369681_1_limnodrilus_hoffmeisterivi_cnh6 mt450685_1_limnodrilus_hoffmeisteri_cw0106 ky369573_1_limnodrilus_hoffmeisterivii_ce8570 ln999086_1_limnodrilus_hoffmeisteri391_t17 ln999319_1_limnodrilus_hoffmeisteri612_t17 ky369625_1_limnodrilus_hoffmeisteriix_ce22808 ky369546_1_limnodrilus_hoffmeisterivii_ce4919 ky369537_1_limnodrilus_hoffmeisteriiv_ce3744 ky369497_1_limnodrilus_hoffmeisterivii_ce2720 ky369735_1_limnodrilus_hoffmeisterivi_cnyd1 ln810408_1_limnodrilus_hoffmeisterit17_132 ky369683_1_limnodrilus_hoffmeisterivi_cnh9 ln999312_1_limnodrilus_hoffmeisteri604_t21 lt905374_1_limnodrilus_hoffmeisteri764_t18 ln999118_1_limnodrilus_hoffmeisteri409_t17 ln999066_1_limnodrilus_hoffmeisteri357_t17 ky369483_1_limnodrilus_hoffmeisteriix_ce1845 ky369468_1_limnodrilus_hoffmeisterii_ce1156 lt905386_1_limnodrilus_hoffmeisteri777_t17 ky369728_1_limnodrilus_hoffmeisterivi_cnwq14 ky369746_1_limnodrilus_hoffmeisterivi_cnyp1 ln999261_1_limnodrilus_hoffmeisteri554_t17 ln810313_1_limnodrilus_hoffmeisterit20_272 ln999180_1_limnodrilus_hoffmeisteri473_t17 ln999211_1_limnodrilus_hoffmeisteri507_t17 ky369495_1_limnodrilus_hoffmeisterivii_ce2712 ln999076_1_limnodrilus_hoffmeisteri371_t17 mf125273_1_limnodrilus_hoffmeisteri ln999225_1_limnodrilus_hoffmeisteri522_t17 ln999359_1_limnodrilus_hoffmeisteri661_t21 ky369741_1_limnodrilus_hoffmeisterivi_cnyd8 lt905405_1_limnodrilus_hoffmeisteri867_t17 ky369472_1_limnodrilus_hoffmeisterii_ce1171 ky369557_1_limnodrilus_hoffmeisterivii_ce5859 ln999316_1_limnodrilus_hoffmeisteri_608_t17 ky369501_1_limnodrilus_hoffmeisteriviii_ce2740 ky369754_1_limnodrilus_hoffmeisterivi_cnyp25 ln999311_1_limnodrilus_hoffmeisteri603_t21 lt905380_1_limnodrilus_hoffmeisteri770_t17 ln810392_1_limnodrilus_hoffmeisterit17_129 ky369699_1_limnodrilus_hoffmeisteriix_cnj3 ln810405_1_limnodrilus_hoffmeisterit17_275 ky369651_1_limnodrilus_hoffmeisteriii_ce22903 ln999191_1_limnodrilus_hoffmeisteri485_t17 ln999345_1_limnodrilus_hoffmeisteri643_t18 ln999070_1_limnodrilus_hoffmeisteri365_t21 ky369541_1_limnodrilus_hoffmeisterix_ce3931 ln999229_1_limnodrilus_hoffmeisteri528_t17 ky369496_1_limnodrilus_hoffmeisterivii_ce2713 ln999354_1_limnodrilus_hoffmeisteri653_t17 ky369630_1_limnodrilus_hoffmeisteriii_ce22875 ky369753_1_limnodrilus_hoffmeisterivi_cnyp9 ky369624_1_limnodrilus_hoffmeisteriix_ce22807 ky369479_1_limnodrilus_hoffmeisteriix_ce1841 ln999314_1_limnodrilus_hoffmeisteri606_t18 lr025066_1_limnodrilus_hoffmeisteri1123_t17 ky369748_1_limnodrilus_hoffmeisterivi_cnyp4 ky369720_1_limnodrilus_hoffmeisterivi_cnl61 ln810314_1_limnodrilus_hoffmeisterit19_269 ln999327_1_limnodrilus_hoffmeisteri620_t18 ky369509_1_limnodrilus_hoffmeisterix_ce2886 ln999082_1_limnodrilus_hoffmeisteri381_t17 ky369540_1_limnodrilus_hoffmeisteriiii_ce3779 ky369680_1_limnodrilus_hoffmeisterivi_cnh5 lt905379_1_limnodrilus_hoffmeisteri769_t18 ln810308_1_limnodrilus_hoffmeisterit21_267 ky369629_1_limnodrilus_hoffmeisteriii_ce22874 ln999268_1_limnodrilus_hoffmeisteri562_t17 ef089357_1_limnodrilus_hoffmeisteri ln999326_1_limnodrilus_hoffmeisteri619_t17 ln810396_1_limnodrilus_hoffmeisterit17_304 ky369477_1_limnodrilus_hoffmeisteriix_ce1784 mt324115_1_limnodrilus_hoffmeisteri_bioug43944-c04 ln999177_1_limnodrilus_hoffmeisteri470_t17 mt450680_1_limnodrilus_hoffmeisteri_cw0007 ky369588_1_limnodrilus_hoffmeisterii_ce8616 ky369613_1_limnodrilus_hoffmeisterivii_ce12501 ln810399_1_limnodrilus_hoffmeisterit17_327 ky369478_1_limnodrilus_hoffmeisteriix_ce1840 ky369703_1_limnodrilus_hoffmeisteriix_cnj8 ky369737_1_limnodrilus_hoffmeisterivi_cnyd3 9450_a003_limnodrilus_hoffmeisteri ky369749_1_limnodrilus_hoffmeisterivi_cnyp5 ky369641_1_limnodrilus_hoffmeisteriii_ce22891 ky369506_1_limnodrilus_hoffmeisterivii_ce2760 ky369591_1_limnodrilus_hoffmeisterix_ce9019 ky369634_1_limnodrilus_hoffmeisteriii_ce22881 ln999101_1_limnodrilus_hoffmeisteri385_t17 lt903826_1_limnodrilus_hoffmeisteri972_t17 ln810410_1_limnodrilus_hoffmeisterit18_342 ky369661_1_limnodrilus_hoffmeisterivii_ce22916 lt905378_1_limnodrilus_hoffmeisteri768_t18 ky369635_1_limnodrilus_hoffmeisterix_ce22883 ky369602_1_limnodrilus_hoffmeisteriix_ce10776 ky369717_1_limnodrilus_hoffmeisterivi_cnl48 ky369730_1_limnodrilus_hoffmeisterivi_cnwq16 lt903793_1_limnodrilus_hoffmeisteri990_t18 ln999337_1_limnodrilus_hoffmeisteri633_t17 ky369646_1_limnodrilus_hoffmeisteriii_ce22897 lt905367_1_limnodrilus_hoffmeisteri754_t17 lt598635_1_limnodrilus_hoffmeisteri935 ln810402_1_limnodrilus_hoffmeisterit17_326 ln810390_1_limnodrilus_hoffmeisterit17_166 ky369544_1_limnodrilus_hoffmeisterix_ce4673 ln999342_1_limnodrilus_hoffmeisteri640_t17 ln810310_1_limnodrilus_hoffmeisterit20_259 ky369529_1_limnodrilus_hoffmeisterii_ce3238 ky369574_1_limnodrilus_hoffmeisterii_ce8591 ln999231_1_limnodrilus_hoffmeisteri530_t17 ln810412_1_limnodrilus_hoffmeisterit18_266 ky369695_1_limnodrilus_hoffmeisterivi_cnh39 ln810401_1_limnodrilus_hoffmeisterit17_134 ky369671_1_limnodrilus_hoffmeisteriix_cnas18 ln999336_1_limnodrilus_hoffmeisteri631_t21 ky369599_1_limnodrilus_hoffmeisterix_ce10649 ln999315_1_limnodrilus_hoffmeisteri607_t21 ky369484_1_limnodrilus_hoffmeisterix_ce1990 ln999084_1_limnodrilus_hoffmeisteri384_t17 ky369523_1_limnodrilus_hoffmeisterivii_ce3120 ln810309_1_limnodrilus_hoffmeisterit20_142 ln999120_1_limnodrilus_hoffmeisteri413_t17 mt450679_1_limnodrilus_hoffmeisteri_cw0006 ky369638_1_limnodrilus_hoffmeisteriii_ce22887 ky369664_1_limnodrilus_hoffmeisterix_cnas4 ky369480_1_limnodrilus_hoffmeisteriix_ce1842 ky369516_1_limnodrilus_hoffmeisteriix_ce3089 ln999119_1_limnodrilus_hoffmeisteri412_t21 ky369740_1_limnodrilus_hoffmeisterivi_cnyd7 ln999376_1_limnodrilus_hoffmeisteri632_t18 ky369672_1_limnodrilus_hoffmeisteriix_cnas23 mt450686_1_limnodrilus_hoffmeisteri_cw0220 ky369712_1_limnodrilus_hoffmeisterivi_cnl1 ln999158_1_limnodrilus_hoffmeisteri450_t17 ky369545_1_limnodrilus_hoffmeisterix_ce4674 ky369747_1_limnodrilus_hoffmeisterivi_cnyp2 eu311398_1_limnodrilus_hoffmeisteri_lh17 ln999161_1_limnodrilus_hoffmeisteri455_t17 ln999361_1_limnodrilus_hoffmeisteri663_t21 ln999133_1_limnodrilus_hoffmeisteri427_t17 ky369571_1_limnodrilus_hoffmeisterivii_ce8552 ky369568_1_limnodrilus_hoffmeisterix_ce7157 ky369645_1_limnodrilus_hoffmeisteriii_ce22896 9450_pl04_limnodrilus_hoffmeisteri ln999280_1_limnodrilus_hoffmeisteri598_t18 ky369631_1_limnodrilus_hoffmeisterivii_ce22877 mn294955_1_limnodrilus_hoffmeisteri_ban_c9 ln999346_1_limnodrilus_hoffmeisteri644_t17 ky369542_1_limnodrilus_hoffmeisterix_ce3932 ky369486_1_limnodrilus_hoffmeisterix_ce1992 ky369507_1_limnodrilus_hoffmeisterivii_ce2761 ky369504_1_limnodrilus_hoffmeisteriviii_ce2743 ky369535_1_limnodrilus_hoffmeisteriiv_ce3738 ky369697_1_limnodrilus_hoffmeisterivi_cnh42 ln999308_1_limnodrilus_hoffmeisteri600_t21 ln999186_1_limnodrilus_hoffmeisteri480_t17 ky369674_1_limnodrilus_hoffmeisterivi_cnf2 ky369656_1_limnodrilus_hoffmeisteriii_ce22908 ky369669_1_limnodrilus_hoffmeisterix_cnas13 ky369618_1_limnodrilus_hoffmeisterix_ce14191 ln999318_1_limnodrilus_hoffmeisteri610_t17 om323983_1_limnodrilus_hoffmeisteriiii_bsp1 ky369726_1_limnodrilus_hoffmeisterivi_cnwq12 ln810388_1_limnodrilus_hoffmeisterit17_172 ky369582_1_limnodrilus_hoffmeisterivii_ce8601 ln999162_1_limnodrilus_hoffmeisteri456_t17 ky369508_1_limnodrilus_hoffmeisterivii_ce2809 ky369663_1_limnodrilus_hoffmeisterix_cnas2 ln810414_1_limnodrilus_hoffmeisterit18_169 ln810403_1_limnodrilus_hoffmeisterit17_325 ky369704_1_limnodrilus_hoffmeisteri_cnj16 ln999067_1_limnodrilus_hoffmeisteri362_t19 ky369742_1_limnodrilus_hoffmeisterivi_cnyd9 ky369464_1_limnodrilus_hoffmeisteriviii_ce608 ky369526_1_limnodrilus_hoffmeisterix_ce3124 ky369556_1_limnodrilus_hoffmeisterivii_ce5858 ky369597_1_limnodrilus_hoffmeisterix_ce10616 ln999370_1_limnodrilus_hoffmeisteri675_t21 ky369583_1_limnodrilus_hoffmeisterii_ce8602 ky369600_1_limnodrilus_hoffmeisteriix_ce10774 ky369532_1_limnodrilus_hoffmeisterii_ce3241 ky369739_1_limnodrilus_hoffmeisterivi_cnyd5 ky369520_1_limnodrilus_hoffmeisteriix_ce3093 ky369489_1_limnodrilus_hoffmeisteriii_ce2281 mt450683_1_limnodrilus_hoffmeisteri_cw0021 hm460271_1_limnodrilus_hoffmeisteri_ce290 ky369521_1_limnodrilus_hoffmeisteriix_ce3094 ky369729_1_limnodrilus_hoffmeisterivi_cnwq15 mk393321_1_limnodrilus_hoffmeisteri_tsjm461 ky369662_1_limnodrilus_hoffmeisteriix_cnas1 ky369610_1_limnodrilus_hoffmeisteriix_ce12493 ln999160_1_limnodrilus_hoffmeisteri454_t17 9450_pl02_limnodrilus_hoffmeisteri ky369576_1_limnodrilus_hoffmeisterii_ce8593 ln999071_1_limnodrilus_hoffmeisteri366_t17 ln999072_1_limnodrilus_hoffmeisteri367_t21 lt903786_1_limnodrilus_hoffmeisteri1017_t17 ky369503_1_limnodrilus_hoffmeisteriviii_ce2742 ky369513_1_limnodrilus_hoffmeisterivii_ce3086 ln999351_1_limnodrilus_hoffmeisteri650_t21 ky369611_1_limnodrilus_hoffmeisteriix_ce12498 ln999210_1_limnodrilus_hoffmeisteri506_t17 ln999065_1_limnodrilus_hoffmeisteri356_t21 ky369575_1_limnodrilus_hoffmeisterivii_ce8592 ky369639_1_limnodrilus_hoffmeisteriii_ce22888 ky369492_1_limnodrilus_hoffmeisterivii_ce2667 ky369636_1_limnodrilus_hoffmeisteriii_ce22884 ln999320_1_limnodrilus_hoffmeisteri613_t20 eu311397_1_limnodrilus_hoffmeisteri_lh2 ln999322_1_limnodrilus_hoffmeisteri615_t21 mn961663_1_limnodrilus_hoffmeisteri_wa12 ky369667_1_limnodrilus_hoffmeisterix_cnas9 ln999159_1_limnodrilus_hoffmeisteri453_t21 ky369593_1_limnodrilus_hoffmeisterivii_ce9827 ln999248_1_limnodrilus_hoffmeisteri541_t17 ky369552_1_limnodrilus_hoffmeisteriiii_ce5201 ky369584_1_limnodrilus_hoffmeisterii_ce8603 ln810413_1_limnodrilus_hoffmeisterit18_322 ky369522_1_limnodrilus_hoffmeisterivii_ce3119 ln999093_1_limnodrilus_hoffmeisteri359_t21 ln999309_1_limnodrilus_hoffmeisteri601_t17 ky369596_1_limnodrilus_hoffmeisterix_ce10615 ky369718_1_limnodrilus_hoffmeisterivi_cnl57 ky369510_1_limnodrilus_hoffmeisterivii_ce2977 ky369725_1_limnodrilus_hoffmeisterivi_cnwq11 ln999103_1_limnodrilus_hoffmeisteri387_t21 ln999127_1_limnodrilus_hoffmeisteri420_t19 ky369731_1_limnodrilus_hoffmeisterivi_cnwq17 ln999131_1_limnodrilus_hoffmeisteri425_t19 ln810394_1_limnodrilus_hoffmeisterit17_128 ky369675_1_limnodrilus_hoffmeisterivi_cnf3 ky369751_1_limnodrilus_hoffmeisterivi_cnyp7 ky369493_1_limnodrilus_hoffmeisteriviii_ce2710 ky369688_1_limnodrilus_hoffmeisterivi_cnh21 ky369716_1_limnodrilus_hoffmeisterivi_cnl17 ky369598_1_limnodrilus_hoffmeisterivii_ce10619 ky369527_1_limnodrilus_hoffmeisterivii_ce3138 mt450681_1_limnodrilus_hoffmeisteri_cw0015 lr025080_1_limnodrilus_hoffmeisteri1144_t17 ln999288_1_limnodrilus_hoffmeisteri577_t18 lt903816_1_limnodrilus_hoffmeisteri941_t18 ky369580_1_limnodrilus_hoffmeisterivii_ce8598 ky369569_1_limnodrilus_hoffmeisterivii_ce7179 ky369653_1_limnodrilus_hoffmeisterivii_ce22905 ky369547_1_limnodrilus_hoffmeisteriix_ce4924 ln999081_1_limnodrilus_hoffmeisteri379_t19 lt899885_1_limnodrilus_hoffmeisteri792_t18 ky369744_1_limnodrilus_hoffmeisterivi_cnyd12 ky369549_1_limnodrilus_hoffmeisterivii_ce5198 ln999112_1_limnodrilus_hoffmeisteri400_t17 ky369538_1_limnodrilus_hoffmeisteriiv_ce3763 ky369690_1_limnodrilus_hoffmeisterivi_cnh27 lt899898_1_limnodrilus_hoffmeisteri750_t16 ky369715_1_limnodrilus_hoffmeisterivi_cnl13 ky369534_1_limnodrilus_hoffmeisteriviii_ce3334 ky369657_1_limnodrilus_hoffmeisteriii_ce22910 ky369732_1_limnodrilus_hoffmeisterivi_cnwq36 lt905370_1_limnodrilus_hoffmeisteri758_t17 ky369586_1_limnodrilus_hoffmeisterii_ce8610 ky369654_1_limnodrilus_hoffmeisterivii_ce22906 ln999080_1_limnodrilus_hoffmeisteri378_t17 ky369691_1_limnodrilus_hoffmeisterivi_cnh29 ky369577_1_limnodrilus_hoffmeisterivii_ce8594 ky369536_1_limnodrilus_hoffmeisteriiii_ce3739 ln999333_1_limnodrilus_hoffmeisteri628_t17 ky369470_1_limnodrilus_hoffmeisterivii_ce1158 ln810406_1_limnodrilus_hoffmeisterit17_274 ky369743_1_limnodrilus_hoffmeisterivi_cnyd11 ky369684_1_limnodrilus_hoffmeisterivi_cnh15 ln999122_1_limnodrilus_hoffmeisteri415_t19 ky369685_1_limnodrilus_hoffmeisterivi_cnh16 ln999115_1_limnodrilus_hoffmeisteri406_t17 ky369482_1_limnodrilus_hoffmeisteriix_ce1844 lt905373_1_limnodrilus_hoffmeisteri763_t18 lt905371_1_limnodrilus_hoffmeisteri759_t17 9450_pl01_limnodrilus_hoffmeisteri ln999242_1_limnodrilus_hoffmeisteri525_t17 ky369539_1_limnodrilus_hoffmeisteriiii_ce3778 mn294954_1_limnodrilus_hoffmeisteri_ban_d9 ky369555_1_limnodrilus_hoffmeisteriiii_ce5204 ky369734_1_limnodrilus_hoffmeisterivii_cnwq42 ln999357_1_limnodrilus_hoffmeisteri656_t17 ky369511_1_limnodrilus_hoffmeisterivii_ce2978 ln810311_1_limnodrilus_hoffmeisterit20_141 ky369487_1_limnodrilus_hoffmeisteriix_ce2126 ky369705_1_limnodrilus_hoffmeisterivii_cnj20 ky369570_1_limnodrilus_hoffmeisteri_ce7448 ln810306_1_limnodrilus_hoffmeisterit21_130 ln999078_1_limnodrilus_hoffmeisteri373_t17 ky369551_1_limnodrilus_hoffmeisterivii_ce5200 ky369473_1_limnodrilus_hoffmeisterii_ce1172 ln999087_1_limnodrilus_hoffmeisteri392_t17 ky369750_1_limnodrilus_hoffmeisterivi_cnyp6 ky369578_1_limnodrilus_hoffmeisterii_ce8595 lt899877_1_limnodrilus_hoffmeisteri858_t18 lr025079_1_limnodrilus_hoffmeisteri1143_t18 ky369553_1_limnodrilus_hoffmeisteriiii_ce5202 ky369558_1_limnodrilus_hoffmeisterivii_ce5860 ky369666_1_limnodrilus_hoffmeisterix_cnas8 ky369649_1_limnodrilus_hoffmeisterivii_ce22900 ky369606_1_limnodrilus_hoffmeisterix_ce10845 lt903809_1_limnodrilus_hoffmeisteri1001_t17 lr025082_1_limnodrilus_hoffmeisteri1147_t17 ky369579_1_limnodrilus_hoffmeisterivii_ce8597 ky369514_1_limnodrilus_hoffmeisteriix_ce3087 ky369601_1_limnodrilus_hoffmeisteriix_ce10775 ln999321_1_limnodrilus_hoffmeisteri614_t21 lr025078_1_limnodrilus_hoffmeisteri1142_t17 ky369612_1_limnodrilus_hoffmeisteriix_ce12499 ky369713_1_limnodrilus_hoffmeisterivi_cnl7 ky369561_1_limnodrilus_hoffmeisterix_ce6575 ln999279_1_limnodrilus_hoffmeisteri597_t18 ef089358_1_limnodrilus_hoffmeisteri ln999107_1_limnodrilus_hoffmeisteri395_t19 ln810398_1_limnodrilus_hoffmeisterit17_332 ln999207_1_limnodrilus_hoffmeisteri504_t17 ky369615_1_limnodrilus_hoffmeisteriviii_ce14181 ky369660_1_limnodrilus_hoffmeisterivii_ce22914 ky369517_1_limnodrilus_hoffmeisteriix_ce3090 ln999221_1_limnodrilus_hoffmeisteri518_t17 ln999094_1_limnodrilus_hoffmeisteri358_t19 ky369467_1_limnodrilus_hoffmeisterii_ce1137 ky369623_1_limnodrilus_hoffmeisteriix_ce22806 ky369668_1_limnodrilus_hoffmeisterix_cnas12 ky369617_1_limnodrilus_hoffmeisteriix_ce14187 ky369594_1_limnodrilus_hoffmeisterix_ce10613 ky369465_1_limnodrilus_hoffmeisteriix_ce778 0.81 0.95 1 0.47 0.75 1 0.8 0.83 0.98 0.96 0.99 0.97 0.45 0.89 0.99 0.99 1 0.51 0.76 1 1 advances in oceanography and limnology advances in oceanography and limnology advances in oceanography and limnology advances in oceanography and limnology advances in oceanography and limnology advances in oceanography and limnology advances in oceanography and limnology advances in oceanography and limnology advances in oceanography and limnology advances in oceanography and limnology advances in oceanography and limnology advances in oceanography and limnology advances in oceanography and limnology advances in oceanography and limnology advances in oceanography and limnology advances in oceanography and limnology advances in oceanography and limnology advances in oceanography and limnology advances in oceanography and limnology advances in oceanography and limnology advances in oceanography and limnology advances in oceanography and limnology advances in oceanography and limnology advances in oceanography and limnology advances in oceanography and limnology advances in oceanography and limnology advances in oceanography and limnology advances in oceanography and limnology layout 1 introduction while in marked decline worldwide, seagrasses are important structural species considered to be “ecosystem engineers”, hosting rich and abundant benthic assemblages and often representing a source of biodiversity for adjacent habitats (e.g. surugiu et al., 2021; barnes et al., 2022; garcía-gómez et article macrobenthos of lagoon ecosystems: a comparison in vegetated and bare sediments paolo magni1,*, maria flavia gravina1,2,3 1consiglio nazionale delle ricerche, istituto per lo studio degli impatti antropici e sostenibilità in ambiente marino (cnr-ias), località sa mardini, torregrande, oristano, italy; 2dipartimento di biologia, università di roma “tor vergata”, rome, italy; 3consorzio nazionale interuniversitario per le scienze del mare (conisma), rome, italy abstract the classic paradigm, not always unequivocal though, that seagrass macrobenthic abundance and diversity are greater than those in adjacent unvegetated areas, was tested in a mediterranean lagoon for which evidence is lacking. we compared the community structure and species composition of macrobenthic assemblages in three distinct areas of the mistras lagoon (sardinia, italy) dominated by i) the seagrass cymodocea nodosa (cym), ii) mixed-macrophyte/detritus (mix), and iii) unvegetated sediments (unv). samplings were conducted in each area twice in spring (april and may) and twice in autumn (october and november) 2010. multivariate analyses showed significant differences among the three areas and a marked separation between the two vegetated (cym and mix) and the unvegetated (unv) areas. the top discriminating species, indicated by simper analysis, were characterized by direct development lacking free living larvae. they included marine species, e.g. cerithium lividulum, microdeutopus gryllotalpa, loripes orbiculatus and gammarus aequicauda, at cym vs mix, whereas the pair discrimination of the areas cym vs unv and mix vs unv mostly depended on the marine abra tenuis, the opportunistic chironomidae and capitella capitata, and the brackish hydrobia acuta. both vegetated areas showed a higher species number than the unvegetated area. differently, abundance was higher in unv than in cym and mix in all sampling dates, most notably in spring likely because of reproductive events, due to both opportunistic (c. capitata, chironomids) and halolimnobic (h. acuta, cerastoderma glaucum) species. overall, the present study demonstrated the importance of biological factors in structuring the macrobenthic assemblages of the mistras lagoon. habitatforming phanerogams hosted species-rich assemblages, species-specific reproductive cycles and adaptive strategies contributed to drive species colonization and abundance distribution. both vegetated and unvegetated patches were found to contribute greatly to the local-scale heterogeneity of the habitat, highlighting the importance of coastal lagoons as hotspots for benthic biodiversity. for these reasons, lagoons must be considered major targets for conservation measures. corresponding author: paolo magni, consiglio nazionale delle ricerche, istituto per lo studio degli impatti antropici e sostenibilità in ambiente marino (cnr-ias), località sa mardini, torregrande, 09170, oristano, italy. e-mail: paolo.magni@cnr.it key words: seagrass; benthic macroinvertebrates; biodiversity, coastal lagoons. authors’ contributions: all the authors have made a substantive intellectual contribution. all the authors have read and approved the final version of the manuscript and agreed to be held accountable for all aspects of the work. conflict of interest: the authors declare no potential conflict of interest. funding: this publication is a result of marbefes (marine biodiversity and ecosystem functioning leading to ecosystem services) project, funded by the european union under the horizon europe programme, “horizon-cl6-2021-biodiv-01” theme, grant agreement no. 101060937 [marbefes.eu]. availability of data and materials: all data generated or analyzed during this study are included in this published article. acknowledgments: we thank giovanni fenzi and andrea satta for their work in the field and in the laboratory and messrs manca of the cooperativa pescatori e molluschicoltori of mistras lagoon for permission to conduct research on the mistras lagoon system. we gratefully acknowledge two anonymous reviewers for providing valuable comments to the original manuscript. received: 29 december 2022. accepted: 8 march 2023. publisher’s note: all claims expressed in this article are solely those of the authors and do not necessarily represent those of their affiliated organizations, or those of the publisher, the editors and the reviewers. any product that may be evaluated in this article or claim that may be made by its manufacturer is not guaranteed or endorsed by the publisher. ©copyright: the author(s), 2023 licensee pagepress, italy advances in oceanography and limnology, 2023; 14:11124 doi: 10.4081/aiol.2023.11124 this work is licensed under a creative commons attribution noncommercial 4.0 international license (cc by-nc 4.0). no nco mm er cia l u se on ly macrobenthos of lagoon ecosystems: a comparison in vegetated and bare sediments 11 al., 2022). in intertidal and subtidal estuarine systems, several studies have been conducted on biodiversity differentials between seagrass and bare sediments, highlighting the prominent role of seagrass in enhancing the benthic biodiversity (barnes and barnes, 2012; barnes, 2013a; surugiu et al., 2021). other studies have compared macrobenthic assemblages in different seagrass species (leopardas et al., 2014; barnes, 2020a, 2020b;), confirming the importance of the architectural complexity of seagrass meadows in supporting higher abundance, species richness and diversity. this said, it should also be acknowledged that the classic paradigm that seagrass supports macrobenthic assemblages with different composition and with greater abundance and diversity with respect to adjacent areas without this cover is not always unequivocal (see nakaoka, 2005; barnes, 2022 and references therein). for instance, departure from this paradigm is found in some intertidal and estuarine systems in relation to local environmental conditions and gradients, assemblage composition and biomass, predatory behaviour and ontogenetic diet shift, latitudinal gradients, rhizome biomass and seasonal cycle of macrophyte vegetation, etc. (see barnes and barnes, 2014; włodarska-kowalczuk et al., 2014; magni et al., 2017 and references therein). these studies highlight the complexity of animal-plant interaction in seagrass ecology, which requires a detailed analysis of the local environmental peculiarities. most importantly, this suggests that both vegetated and unvegetated areas need to be considered mutually for conservation purposes to prevent the potential loss of local biodiversity and to understand the ecological processes structuring the benthic assemblages in these complex habitats (barnes, 2013b; barnes and hamylton, 2016; barnes 2021). it has been shown that in coastal lagoons continuous stands of seagrass offer stability to well structured and taxa-rich macrozoobenthic assemblages (sfriso et al., 2001; tagliapietra et al., 2016). this can also be due to a series of additional factors, such as the shading effect and protection of seagrass meadows from high temperature in summer, especially in shallow waters, and a better oxygenation of the water and sediments in the presence of the seagrass (tagliapietra et al., 2016). moving away from the meadow to the bare sediment, the temporal fluctuations of the community structure become wider and the assemblages reduce their resistance and resilience, due to worsening or changing environmental conditions, such as increased sediment organic enrichment (magni et al., 2015; leopardas et al., 2016) or reduced structural complexity of the habitat (bachelet et al., 2000; leopardas et al., 2014), respectively. however, very few studies looking at seagrass vs bare sediment benthic biodiversity are available in coastal lagoons (magni et al., 2017; barnes, 2022) and, to our knowledge, no studies have compared the community structure of macrobenthic assemblages in multiple habitats, including vegetated and non-vegetated areas, within a mediterranean lagoon system. the present study was carried out in one of these mediterranean lagoons, the mistras lagoon, located along the western coast of sardinia. we aimed to evaluate the differences in macrobenthic assemblages between seagrass-dominated and unvegetated seabeds by analyzing and comparing the community structure and species composition of macrobenthic assemblages in three distinct areas, within the same lagoon, characterized by the dominance of i) the seagrass cymodocea nodosa, ii) mixed-macrophyte/detritus, and iii) unvegetated sediments. materials and methods the mistras lagoon the mistras lagoon is a semi-enclosed coastal barrier system (~4.5 km2, mean depth <1 m) located on the western coast of sardinia island (italy, western mediterranean sea) (figure 1). connected to the gulf of oristano through its mouth, it is part of the so-called oristano lagoon-gulf (olg) system which comprises the gulf of oristano (150 km2; maximum depth 24 m), several salt marshes and the lagoons of cabras, santa giusta, s’ena arrubia, corru s’ittiri and san giovannimarceddì (magni et al., 2008a). these lagoons are characterized by eutrophic conditions, with the dominance of small sized phytoplankton species and the occurrence of harmful algal species (e.g. cabras lagoon) or the major presence of macrophyte (e.g. s’ena arrubia and santa giusta lagoons) (padedda et al., 2012; satta et al., 2014; pulina et al., 2018). sediments are generally organically enriched, leading to impoverished benthic communities (magni et al., 2005; brundu and magni, 2021) and occasional dystrophic events (murenu et al. 2004; magni et al., 2008b). on the contrary, the mistras lagoon is characterized by oligotrophic waters (sechi, 1982; specchiulli et al., 2018) and sandy sediments with a low content of organic matter (om), especially in the unvegetated areas, owing to negligible riverine discharge and low sediment supply (pascucci et al., 2018). major pollution sources are absent, leading a low-pressure index compared to other sardinian lagoons (gravina et al., 2020). during dry periods with high evaporation rates in summer, satellite ponds and natural salt flats are formed in shallower and confined areas where salinity can exceed 50 psu. fishing activity in the mistras lagoon is carried extensively using a capture system made by workers and fish traps named “bertovelli”. the mistras lagoon is listed in the ramsar convention on wetlands (d.m. 04/03/82) and is a special protection area under the nature 2000 network (zps itb034006, sic itb030034), thus subjected to local protection measures. in the lagoon there are meadows of various species of aquatic phanerogams, including the dominant c. nodosa followed by zostera noltii and ruppia maritima, and several macroalgal species, such as gracilaria sp., ulva sp. and enteromorpha sp., (camarda, 1995). the lagoon is surrounded by luxuriant glasswort dominated by salicornia fruticosa, with the presence of obione portulacoides, salsola soda, limonium vulgare, aeluropus litoralis, arthrocnemum glaucum, atriplex sp. and scirpus sp. (fenzi, 2013). study areas, sampling strategy and sample treatment three different sampling areas were selected based on the morphology of the lagoon and the presence and distribution of vegetation. in particular, the seagrass area (hereafter, cym) was located in the southwest sector of the lagoon and was dominated by c. nodosa meadows, with scattered patches of z. noltii. the no nco mm er cia l u se on ly p. magni and m.f. gravina12 second vegetated area, named mixed (hereafter, mix), was located closer to the lagoon’s mouth and was characterized by sparse patches of both c. nodosa and z. noltii, the red alga gracilaria sp. developing from autumn to spring, and vegetation-derived detritus. the third area was unvegetated (hereafter, unv) and was located in the north-eastern and most confined sector of the lagoon. this area was characterized by sandy sediments and low om content, without vegetation except sporadic patches of r. maritima and enteromorpha sp. in spring (figure 1). sampling was conducted in each of the three selected areas, i.e. cym, mix and unv, twice in spring, i.e. april and may, and twice in autumn, i.e. october and november, 2010. in each area and each date, 3 stations were randomly selected 10’s meters apart and at each station four replicate samples, meters apart, were randomly collected using a box-corer (170 cm2, penetration depth 15 cm), for a total of 144 samples. each sample was sieved on a 0.5 mm mesh-size and the residue was preserved in a 4% buffered formaldehyde solution for the determination of macrozoobenthos. in the laboratory, macrobenthic organisms from each sample were sorted, counted and identified to the species level whenever possible using a stereomicroscope. data analysis macrozoobenthic data were analyzed by means of ordination technique non-metric multidimensional scaling (nmds) using the mean abundance of four replicates for each of the three stations in each area and each sampling date based on (dis)similarity bray-curtis measure. significance tests for differences among spatial and temporal replicates within and between areas were performed by means of the anosim test. the simper test was carried out in order to evaluate the taxa contributing most to (dis)similarity between and within temporal and spatial groups of replicates. in order to analyze spatiotemporal changes in the macrobenthic assemblages, the following biotic variables were computed: species richness (number of species), total abundance (n, number of individuals m-2), diversity (h’), calculated by the shannon-wiener index, and evenness by the pielou index. significance of the biotic variables was tested by means of a two-way permanova. results macrobenthic assemblages in the present study, 6524 individuals belonging to 33 taxa were classified, of which 39.9 % were molluscs, followed by gastropods (16.1%), crustaceans (15,5%) and polychaetes (14.9%). the ten most abundant taxa, accounting for 88.8% of the total abundance, included the bivalves abra tenuis (26.3%), cerastoderma glaucum (6.6%) and loripes orbiculatus (4.6%), the gastropods cerithium lividulum (8.6%) and hydrobia acuta (7.6%), the polychaetes capitella capitata (7.7%) and perinereis rullieri (3.9%), the amphipods microdeutopus gryllotalpa (6.5%) and gammarus aequicauda (4.4%), and chironomidae (12.7%). figure 1. location of the mistras lagoon and the three study areas: cymodocea nodosa-dominated (cym) area, mixed-macrophyte/detritus (mix) area, and unvegetated sediment (unv) area. no nco mm er cia l u se on ly macrobenthos of lagoon ecosystems: a comparison in vegetated and bare sediments 13 multivariate analyses the one-way anosim test revealed significant differences among the three areas and a marked separation between the two vegetated (cym and mix) and the unvegetated (unv) areas, with a global r=0.74, and a within and between mean ranks of 156.3 and 388.5 (p=0.0001), respectively. the nmds ordination model confirmed this separation, with unv being most isolated irrespective of seasonal changes. there was instead some overlap between cym and mix as a result of a high similarity between the two biocoenoses (figure 2). furthermore, on a seasonal basis, the two autumn samplings at unv were more dispersed than those at cym and mix, due to a major reduction of both abundance and species richness and diversity. the percentage contribution to the average dissimilarity between the sampling areas, measured by the simper analysis, was reported for the top discriminating species accounting for up to 95% of the cumulative percentage (table 1). these latter included marine species, e.g. c. lividulum, m. gryllotalpa, l. orbiculatus and g. aequicauda, at cym vs mix, whereas the pair discrimination of the areas cym vs unv, and mix vs unv mostly depended on the marine species abra tenuis, the opportunistic taxa, e.g. chironomidae and c. capitata, and the brackish species h. acuta. two-way permanova conducted on the biotic variables showed significant differences in the species richness, abundance, diversity and evenness between areas and months (table 2). figure 2. non-metric multidimensional scaling (nmds) ordination plot of macrozoobenthic abundance data for the three investigated areas of the mistras lagoon: white symbols refer to cym, gray to mix and black to unv. circles indicate april, inverted triangles may, squares october and triangles november (each symbol represents the mean of 4 replicates, see m&m) (stress=0.1). figure 3. mean values (n=12 ± standard error, se) of the total abundance of individuals (n. ind. 170 cm–2), species richness (s), shannon-wiener diversity index and evenness pielou index in the three study areas in the mistras lagoon. a, april; m, may; o, october; n, november. no nco mm er cia l u se on ly p. magni and m.f. gravina14 table 1. list of the top discriminating taxa with contribution more than 5% by simper analysis at the three investigated areas, c. nodosa-dominated (cym), mixed-macrophyte/detritus (mix) and unvegetated (unv). species category: o, opportunistic; c, common in coastal sheltered waters; b, brackish; m, properly marine. cym vs mix average contribution cumulative average average dissimilarity % % abundance abundance cym mix cerithium lividulum m 8.39 13.71 13.71 7.77 3.88 microdeutopus gryllotalpa c 8.36 13.67 27.37 6.46 2.23 loripes orbiculatus c 6.98 11.41 38.78 3.83 2.40 gammarus aequicauda c 6.50 10.62 49.41 3.50 2.33 abra tenuis m 4.74 7.74 57.15 1.63 1.67 sphaeroma serratum m 4.60 7.52 64.67 4.00 0.25 nainereis laevigata c 3.93 6.42 71.08 2.19 0.56 perinereis rullieri m 3.66 5.98 77.06 2.06 3.29 abra segmentum c 3.59 5.86 82.92 1.08 2.15 chironomidae o 2.11 3.45 86.37 0.19 1.04 tanais dulongii o 1.33 2.17 88.54 1.15 0.06 cyathura carinata b 0.94 1.53 90.07 0.73 0.00 gibbula adansoni m 0.85 1.39 91.46 0.08 0.42 cerastoderma glaucum b 0.79 1.30 92.76 0.23 0.35 phylo norvegicus c 0.61 1.00 93.76 0.65 0.02 dispio uncinata m 0.56 0.91 94.67 0.00 0.25 pirenella conica b 0.51 0.84 95.51 0.00 0.27 cym vs unv average contribution cumulative average average dissimilarity % % abundance abundance cym mix abra tenuis m 26.91 27.93 27.93 1.63 32.40 chironomidae o 16.39 17.01 44.94 0.19 16.00 capitella capitata o 8.23 8.54 53.48 0.02 10.30 hydrobia acuta b 8.15 8.46 61.94 0.00 10.30 cerithium lividulum m 8.11 8.42 70.36 7.77 0.00 cerastoderma glaucum b 6.08 6.31 76.67 0.23 8.33 microdeutopus gryllotalpa c 5.50 5.71 82.38 6.46 0.13 loripes orbiculatus c 3.17 3.29 85.68 3.83 0.00 sphaeroma serratum m 2.95 3.06 88.74 4.00 0.00 gammarus aequicauda c 2.64 2.74 91.47 3.50 0.17 perinereis rullieri m 2.34 2.43 93.90 2.06 0.00 nainereis laevigata c 2.05 2.13 96.03 2.19 0.00 mix vs unv average contribution cumulative average average dissimilarity % % abundance abundance cym mix abra tenuis m 28.62 30.90 30.90 1.67 32.40 chironomidae o 17.80 19.22 50.13 1.04 16.00 hydrobia acuta b 8.86 9.57 59.70 0.00 10.30 capitella capitata o 8.72 9.41 69.11 0.19 10.30 cerastoderma glaucum b 6.51 7.03 76.14 0.35 8.33 cerithium lividulum m 4.75 5.13 81.27 3.88 0.00 perinereis rullieri m 4.00 4.32 85.59 3.29 0.00 gammarus aequicauda c 2.81 3.03 88.62 2.33 0.17 abra segmentum c 2.72 2.94 91.56 2.15 0.00 microdeutopus gryllotalpa c 2.57 2.78 94.34 2.23 0.13 loripes orbiculatus c 2.32 2.51 96.85 2.40 0.00 no nco mm er cia l u se on ly macrobenthos of lagoon ecosystems: a comparison in vegetated and bare sediments 15 univariate analyses the mean number of species was highest in cym in april (17.3±0.6) and lowest in unv in november (4.0±0.0) (figure 3). overall, both vegetated areas (cym and mix) showed a higher species richness than unv, yet with a progressive reduction from april to november. in particular, species number was significantly higher in spring and autumn in cym, and significantly higher in april than in the other three dates in mix. instead, species number was constantly lowest in unv where it did not show significant differences between dates. similarly, the diversity index h’ was generally higher in cym and mix than in unv which showed no significant differences between dates. differently, the mean abundance was highest in unv in may (144.7±32.5) and lowest in mix in october (11.7±2.5), with a marked reduction in autumn especially in unv. overall, abundance was higher in unv than in cym and mix in all sampling dates. accordingly, evenness was lowest in cym in april and may, and highest in unv in april (0.87±0.08), while mix showed the highest eveness values in may, october and november (figure 3). discussion the present study demonstrated that both seagrass-dominated and mixed/detritus beds hosted a higher macrobenthic species richness and diversity than unvegetated sediments, the latter being characterized by very different assemblages in terms of community structure and species composition. on the contrary, the abundance was considerably higher in the bare sediments than in the vegetated areas. peaks in the number of individuals in unv were due to both opportunistic (capitella capitata, chironomids) and halolimnobic (hydrobia acuta, cerastoderma glaucum) species, which most notably occurred in spring, likely as a result of reproductive events. notwithstanding an overall decline from spring to autumn, the abundance remained higher in unv than in cym and mix, indicating that the unvegetated areas sustain large populations. the lowest evenness values recorded in unv in autumn indicated that variations in the community structure are mainly due to changes in the abundance rather than in the number of species which in turn remained similar over time. similarly, barnes (2021) found that in the knysna estuarine bay (south africa) the endangered dwarf-eelgrass, zostera capensis, supports fewer animals than adjacent unvegetated areas. this reportedly unusual situation was explained by a probable of lack of bioturbators in bare sediments, whose activity may cause a significant reduction of surface and nearsurface animals. in the mistras lagoon, marked differences between areas were also indicated by the seasonal changes of the macrobenthic assemblages. as highlighted by the nmds plot, both spring and autumn replicates in the two vegetated areas were spatially very close one another, with a marked distinction between the two seasons. the observed decline in the number of species and individuals at cym and mix over the months can be explained by the expected seasonal cycles of lagoonal environments (magni et al., 2005; 2015; tagliapietra et al., 2016), including the extent of algae and plant cover providing important food sources, especially in spring, for many benthic species (magni et al., 2008b; como et al., 2012; zheng et al., 2020). differently, the autumn replicates of the unvegetated area appeared to be much more separated one another than the spring ones. the larger spread of the autumn replicates in unv than in cym and mix can be related to a greater suffering and a slower recovery of the benthic assemblage in unv from environmental stress occurring in the mistras lagoon in summer, when a significant increase in temperature and salinity is recorded (unpublished data). on the contrary, the function of seagrass meadows in offering protection from “natural stress” was clearly expressed by the benthic assemblage at cym that displayed greater spatial and temporal stability of abundance and composition (tagliapietra et al., 2016; see also barnes, 2023), and high levels of resilience to the harsh environmental conditions found in summer. in the mistras lagoon, faunal differences in terms of species composition were especially due to several marine species and species typical of coastal sheltered waters (gravina et al., 2020), which were found much more numerous in the two vegetated areas. as indicated by the simper analysis, most of the species that contributed up to the cumulative 95% were amphipods (m. gryllotalpa, g. aequicauda), isopods (s. serratum, c. carinata), tanaids (t. dulongii) and two bivalves, l. lacteus and a. tenuis (see currás and mora, 1996; holmes et al., 2004), all being characterized by direct development lacking free living ta bl e 2. r es ul ts o f t w ow ay p e r m a n o v a a m on g ar ea s an d da te s (p er m ut at io n n : 9 99 9, d f = d eg re es o f f re ed om ). s pe ci es r ic hn es s a bu nd an ce d iv er si ty (h ’) e ve nn es s (p ie lo u) so ur ce s um o f d f m ea n p se ud o p s um o f df m ea np se ud o p s um o f df m ea n p se ud o p s um o f df m ea n p se ud o p s qu ar es sq ua re -f s qu ar es s qu ar e f sq ua re s s qu ar e f sq ua re s s qu ar e -f a re as 1. 00 2 0 .5 0 4 2. 97 0 .0 00 1 1 .5 0 2 0 .7 5 2 2. 23 0 .0 00 1 0 .4 3 2 0. 21 3 2. 87 0. 00 01 0 .0 4 2 0 .0 2 5. 21 0 .0 09 4 d at es 0. 43 3 0 .1 4 1 2. 40 0 .0 00 3 1 .3 8 3 0 .4 6 1 3. 63 0 .0 00 1 0 .2 6 3 0. 09 1 3. 45 0. 00 01 0 .0 4 3 0 .0 1 3. 58 0 .0 23 7 in te ra ct io n 0 .1 8 6 0 .0 3 2. 56 0 .0 38 6 1 .0 8 6 0 .1 8 5 .3 7 0. 00 01 0 .0 9 6 0. 01 2 .2 6 0 .0 58 1 0 .0 5 6 0 .0 1 2. 27 0 .0 61 9 r es id ua l 0. 28 2 4 0 .0 1 0 .8 1 2 4 0. 03 0 .1 6 2 4 0 .0 1 0 .0 9 2 4 0 .0 0 to ta l 1. 89 3 5 4. 76 3 5 0. 93 3 5 0 .2 2 3 5 no nco mm er cia l u se on ly p. magni and m.f. gravina16 larvae. they also included the nonplanktotrophic gastropod c. lividulum which is able of crawling directly on the bottom after hatching (garilli and galletti, 2006; evangelisti et al., 2016). only a few of the main discriminating species revealed by the simper analysis were opportunistic and brackish/halolimnobic species which contributed to characterize the unvegetated area. they comprised chironomids, c. capitata, h. acuta and c. glaucum, the latter two species found >10 times more abundant in the unvegetated area. marked differences between areas were also highlighted by the distribution of two congeneric species, abra segmentum and abra tenuis, which relates to their different biological cycles and larval phases. in particular, a. segmentum, characterized by a long-life pelagic larval phase typical of bivalves, was dominant and exclusive to the two vegetated areas. this suggests that the nearby marine environment can be a source of new recruits, while too confined unvegetated areas may be hardly colonisable by this bivalve. differently, a. tenuis was dominant in very high numbers in the unvegetated area indicating its ability to maintain a considerable population in confined areas due to its direct development which does not require the marine environment as an external donor (holmes et al., 2004). a similar adaptive strategy was shown by opportunistic (c. capitata, chironomids) and halolimnobic (h. acuta, c. glaucum) species which were able to occupy very confined habitats due to their reproduction and adaptation modes. these results show that only a portion of marine species occurring in the mistras lagoon can be sustained by larval supply from the marine environment, while another pool of species lacking direct connectivity with the sea, but characterized by direct development, is able to self-sustain their populations. the results of the present study add new knowledge to the spatial and temporal distribution of benthic assemblages in vegetated coastal environments, contributing to the understanding of differences (or similarities) between seagrass and bare sediments. it follows conspicuous research on macrobenthic seagrass ecology conducted in the intertidal zone (e.g., barnes and barnes, 2012, 2014; barnes, 2020a, 2020b), with fewer comparative studies carried out in coastal marine habitats (włodarska-kowalczuk et al., 2014; surugiu et al., 2021) and coastal lagoons (magni et al., 2017; barnes, 2022; hu et al., 2022). furthermore, although several ecological studies have been carried out on macrophyte-derived detritus (e.g. see review by lepoint and hyndes, 2022), few works have examined the benthic assemblages in leaf litter and respective mother plants in the same environmental settings (ólafsson et al., 2013). one of these studies was conducted outside the mistras lagoon in the gulf of oristano (como et al., 2008). the authors found that two seagrasses, posidonia oceanica and cymodocea nodosa, had similar species composition and differed from a third habitat, i.e. leaf litter beds, for the exclusive presence of hard-bottom species, such as the tunicate phallusia fumigata, and seagrass-associated species, such as the polychaete syllis garciai and the decapod paguristes syrtensis. in contrast, the leaf litter sediments were characterized by the bivalves a. alba and c. glaucum, not found in the seagrasses, and by the bivalves l. lacteus and ruditapes decussatus, and the polychaete prionospio multibranchiata, exclusive to leaf litter beds. conclusions the present study demonstrated the importance of biological factors in structuring the macrobenthos of the mistras lagoon in both vegetated and non-vegetated areas. the presence of habitat-forming phanerogams, such as c. nodosa, z. noltii and ruppia maritima, sustained well-structured and diverse benthic assemblages that were lacking in bare sediments. our results agree with those reported from other lagoons and sheltered coastal areas worldwide (bachelet et al., 2000; sfriso et al. 2001; blanchet et al., 2005; brito et al., 2005; yamada et al., 2007; barnes and barnes, 2012; magni et al., 2015, tagliapietra et al., 2016; magni et al., 2017; surugiu et al., 2021). furthermore, the life-history traits of the species, such as species-specific reproductive cycles and adaptive strategies (nonnis marzano et al., 2010; cardone et al., 2014; gravina et al., 2020), were important drivers for the colonization of benthic species in the confined areas. we anticipate that, in addition to biological factors, environmental constraints, such as confinement, are also to be considered to fully explore the reasons for differences in the benthic community of the mistras lagoon. in particular, the marked isolation of the unvegetated area and the limited water exchange with the sea and other areas of the lagoon, may enhance the “natural stress” due to larger variations in abiotic variables, e.g. salinity. this could affect both the flora and the fauna of the lagoon and will be the focus of a subsequent study. finally, the present study demonstrates that both vegetated and unvegetated patches contribute greatly to the local-scale heterogeneity of the lagoon ecosystems, highlighting the peculiarity and importance of these habitats as hotspots of benthic biodiversity. for these reasons, coastal lagoons must be considered major targets for conservation measures. references bachelet g, de montaudouin x, auby i, labourg pj, 2000. seasonal changes in macrophyte and macrozoobenthos assemblages in three coastal lagoons under varying degrees of eutrophication. ices j. mar. sci. 57:1495-506. barnes rsk, 2013a. distribution patterns of macrobenthic biodiversity in the intertidal seagrass beds of an estuarine system, and their conservation significance. biodivers. conserv. 22:357–72. barnes rsk, 2013b. spatial stability of macrobenthic seagrass biodiversity. mar. ecol. progr. ser. 493:127–39. barnes rsk, 2020a. do different sympatric seagrasses support macrobenthic faunas of differing composition, abundance, biodiversity or patchiness? mar. environ. res. 160:104983. barnes rsk, 2020b. do species display characteristic intraspecific levels of patchiness in a given habitat type? the case of intertidal seagrass macrobenthos. mar. biol. 167:1-12. barnes rsk, 2021. patterns of seagrass macrobenthic biodiversity in the warm-temperate knysna estuarine bay, western cape: a review. aquat. ecol. 55:327–45. barnes rsk, 2022. biodiversity differentials between seagrass and adjacent bare sediment change along an estuarine gradient. estuar. coast. shelf sci. 274:107951. barnes rsk, 2023. seagrass macrobenthic biodiversity does not no nco mm er cia l u se on ly macrobenthos of lagoon ecosystems: a comparison in vegetated and bare sediments 17 vary in conformity with a leaky-lagoonal confinement gradient. mar. environ. res. 185:105897. barnes rsk, barnes mks, 2012. shore height and differentials between macrobenthic assemblages in vegetated and unvegetated areas of an intertidal sandflat. estuar. coast. shelf sci. 106:112–20. barnes rsk, barnes mks, 2014. biodiversity differentials between the numerically dominant macrobenthos of seagrass and adjacent unvegetated sand in the absence of sandflat bioturbation. mar. environ. res. 99:34–43. barnes rsk, claassens l, seath j, 2022. where ecologically ‘tis better to go brown than green: enhanced seagrass macrobenthic biodiversity within the canals of a brownfield coastal marina. biodivers. conserv. 31:2981–97. barnes rsk, hamylton s, 2016. on the very edge: faunal and functional responses to the interface between benthic seagrass and unvegetated-sand assemblages. mar. ecol. prog. ser. 553:33–48. blanchet h, de montaudouin x, chardy p, bachelet g, 2005. structuring factors and recent changes in subtidal macrozoobenthic communities of a coastal lagoon, arcachon bay (france). estuar. coast. shelf sci. 64:561-76. brito mc, martin d, núñez j, 2005. polychaetes associated to a cymodocea nodosa meadow in the canary islands: assemblage structure, temporal variability and vertical distribution compared to other mediterranean seagrass meadows. mar. biol. 146:467–81. brundu g, magni p, 2021. context-dependent effect of serpulid reefs on the variability of soft-bottom macrobenthic assemblages in three mediterranean lagoons (sardinia, italy). estuar. coast. shelf sci. 262:107589. camarda i, 1995. un sistema di aree di interesse botanico per la salvaguardia della biodiversità floristica della sardegna. bollettino della società sarda di scienze naturali, 30:245-95. cardone f, corriero g, fianchini a, et al., 2014. biodiversity of transitional waters: species composition and comparative analysis of hard bottom communities from south-eastern italian coast. j. mar. biol. ass. uk 94:25–34. como s, magni p, baroli m, et al., 2008. comparative analysis of macrofaunal species richness and composition in posidonia oceanica, cymodocea nodosa and leaf litter beds. mar. biol. 153:1087-101. como s, magni p, van der velde g, et al., 2012. spatial variations in δ13c and δ15n values of primary consumers in a coastal lagoon. estuar. coast. shelf sci. 115:300-8. currás, a, mora j, 1996. aspectos de la dinámica de poblaciones de loripes lacteus (linnaeus, 1758) (mollusca, bivalvia) en fondos zostera spp. de la ría del eo (no de españa). nova acta científica compostelana (bioloxía) 6:167-78. evangelisti f, bonfitto a, morassi m, sabelli b, 2016. how many native cerithium species in the mediterranean sea. an integrative taxonomic approach. j. mollusc. stud. 82:292–304. fenzi ga, 2013. variabilità spazio-temporale dei popolamenti macrozoobentonici in relazione ai cambiamenti stagionali in un ambiente lagunare (mistras, sardegna occidentale) con basso impatto antropico ed elevata stabilità ecologica. tesi dottorale, università degli studi della tuscia, viterbo, italy. garcía-gómez g, garcía-herrero á, sánchez n, et al., 2022. meiofauna is an important, yet often overlooked, component of biodiversity in the ecosystem formed by posidonia oceanica. invert. biol. 141:e12377. garilli v, galletti l, 2006. taxonomical characters for distinguishing cerithium lividulum risso, 1826, and c. renovatum monterosato, 1884 (gastropoda, caenogastropoda, cerithiidae). basteria, 70:109-22. gravina mf, cabiddu s, como s, et al, 2020b. disentangling heterogeneity and commonalities in nanotidal mediterranean lagoons through environmental features and macrozoobenthic assemblages. estuar. coast. shelf sci. 237:106688. holmes sp, dekker r, williams di, 2004. population dynamics and genetic differentiation in the bivalve mollusc abra tenuis: aplanic dispersal. mar. ecol. prog. ser. 268:131–40. hu c, liu y, yang x, et al., 2022. functional trait responses of macrobenthic communities in seagrass microhabitats of a temperate lagoon. mar. pollut. bull. 177:113491. leopardas v, honda k, go ga, et al., 2016. variation in macrofaunal communities of sea grass beds along a pollution gradient in bolinao, northwestern philippines. mar. poll. bull. 105:310-8. leopardas v, uy w, nakaoka m, 2014. benthic macrofaunal assemblages in multispecific seagrass meadows of the southern philippines: variation among vegetation dominated by different seagrass species. j. exp. mar. biol. ecol. 457:71-80. lepoint g, hyndes ga, 2022. tropicalization of seagrass macrophytodetritus accumulations and associated food webs. front. mar. sci. 9:943841. magni p, como s, cucco a, et al., 2008a. a multidisciplinary and ecosystemic approach in the oristano lagoon-gulf system (sardinia, italy) as a tool in management plans. transit. water bull. 2:41-62. magni p, como s, kamijo a, montani s, 2017. effects of zostera marina on the patterns of spatial distribution of sediments and macrozoobenthos in the boreal lagoon of furen (hokkaido, japan). mar. environ. res. 131c:90-102. magni p, micheletti, s, casu d, et al., 2005. relationships between chemical characteristics of sediments and macrofaunal communities in the cabras lagoon (western mediterranean, italy). hydrobiologia 550:109-15. magni p, rajagopal s, van der velde g, et al., 2008b. sediment features, macrozoobenthic assemblages and trophic relationships (δ13c and δ15n analysis) following a dystrophic event with anoxia and sulphide development in the santa giusta lagoon (western sardinia, italy). mar. pollut. bull. 57:125-36. murenu m, olita a, sabatini a, follesa mc, cau a, 2004. dystrophy effects on the liza ramada (risso, 1826) (pisces, mugilidae) population in the cabras lagoon (central-western sardinia). chem. ecol. 20(s1):s425–s433. nakaoka m, 2005. plant-animal interactions in seagrass beds: ongoing and future challenges for understanding population and community dynamics. popul. ecol. 47:167–77. nonnis marzano c, gravina mf, fianchini a, et al., 2010. the macrozoobenthos of the karavasta lagoon system (albania): local assemblages and geographical comparisons. mar. ecol. 31:622–32. ólafsson e, aarnio k, bonsdorff e, arroyo nl, 2013. fauna of the green alga cladophora glomerata in the baltic sea: density, diversity, and algal decomposition stage. marine biology 160: 2353–62. no nco mm er cia l u se on ly p. magni and m.f. gravina18 padedda bm, pulina s, magni p, et al., 2012. phytoplankton dynamics in relation to environmental changes in a phytoplankton-dominated mediterranean lagoon (cabras lagoon, italy). adv. limnol oceanogr. 3:147-169. pascucci v, de falco g, del vais c, et al., 2018. climate changes and human impact on the mistras coastal barrier system (w sardinia, italy). mar. geol. 395:271-84. perez m, duarte cm, romero j, et al., 1994. growth plasticity in cymodocea nodosa stands: the importance of nutrient supply. aquat bot 47:249–64. pulina s, satta ct, padedda bm, et al., 2018. seasonal variations of phytoplankton size structure in relation to environmental variables in three mediterranean shallow coastal lagoons. estuar. coast. shelf sci. 212:95-104. ramsar convention, 1971. convention on wetlands of international importance especially as waterfowl habitat. ramsar (iran), 2 february 1971. un treaty series no. 14583. as amended by the paris protocol, 3 december 1982, and regina amendments, 28 may 1987. available from: http://www.ramsar.org/key_conv_e.html. satta ct, anglès s, garcés e, et al, 2014. dinoflagellate cyst assemblages in surface sediments from three shallow mediterranean lagoons (sardinia, north western mediterranean sea). estuar. coast. 37:646-63. sechi n, 1982. lo stato trofico di alcuni stagni salmastri costieri della sardegna. boll. soc. sarda sci. nat. 21:285-95. sfriso a, birkemeyer t, ghetti pf, 2001. benthic macrofauna changes in areas of venice lagoon populated by seagrasses or seaweeds. mar. environ. res. 52:323-49. specchiulli a, cilenti l, d’adamo r, et al, 2018. dissolved organic matter dynamics in mediterranean lagoons: the relationship between doc and cdom. mar. chem. 202:37-48. surugiu v, teacâ a, svedu i, quijón pa, 2021. a hotspot in the romanian black sea: eelgrass beds drive local biodiversity in surrounding bare sediments. front. mar. sci. 8:745137. tagliapietra d, pessa g, cornello m, et al., 2016. temporal distribution of intertidal macrozoobenthic assemblages in a nanozostera noltii-dominated area (lagoon of venice). mar. environ. res. 114:31-9. włodarska-kowalczuk m, jankowska e, kotwicki l, balazy p, 2014. evidence of season-dependency in vegetation effects on macrofauna in temperate seagrass meadows (baltic sea). plos one 9:e100788. yamada k, hori m, tanaka y, et al., 2007. temporal and spatial macrofaunal community changes along a salinity gradient in seagrass meadows of akkeshi-ko estuary and akkeshi bay, northern japan. hydrobiologia 592:345-58. zheng x, como s, huang l, magni p, 2020. temporal changes of a food web structure driven by different primary producers in a subtropical eutrophic lagoon. mar. environ. res. 161:105-28. no nco mm er cia l u se on ly advances in oceanography and limnology layout 1 article introduction mediterranean micro-tidal environments are transitional systems characterized by a tidal range <2m, where salt-wedge intrusion along the final stretch of rivers plays an important role in shaping biological communities (boothroyd, 1978; davies, 1964; kiriaki et al., 2010; van der tuin, 1991). the extent of this process depends on several geomorphological and physical factors, such as, for example, the river mouth configuration or man-made barrier-like dams (haralambidou et al., 2010; lichter et al., 2011). from a geomorphological point of view, a fennel-shaped river mouth can be defined as an estuary, in which the mixing between seawater and freshwater generates strong salinity gradients and the presence of a typical brackish water layer (perillo, 1995; poulos et al., 1993; valle-levinson, 2010). these systems are particularly vulnerable to climate change, since strong seasonality and variability in precipitation produce large inter-annual fluctuations in river flow and water mass exchange with the sea (gasith and resh, 1999; kennish, 2002, 2021; newman et al., 2005; ustin et al., 2014; verri et al., 2018;). in the last decade, in fact, a reduction of net rainfall of ~33% of the annual mean was reported for southern italian regions, mostly due to a reduced contribution of winter rainfall (caloiero et al., 2021; polemio and casarano 2004, 2008). the effects of the drought period and human interventions like land reclamation, construction of river embankments, and water extraction activities significantly contributed to altering the ecological profiles of these areas (dugdale et al., 2012; hearn and robson, 2001; kennish, 2021; newman et al., 2005; richard et al., 2018;). the combined effect of saltwater intrusion and river discharge produces significant changes in nutrient loads, since the extent to which nutrients are retained in rivers and then made available during the biomass development phytoplankton community as monitoring tool in the terminal stretch of a micro-tidal estuary facing the tyrrhenian sea francesco bolinesi1, emanuela serino1, angelo carotenuto1, silvia fanina1 and olga mangoni1,2 1department of biology, university of naples federico ii, complesso universitario di monte sant’angelo, naples; 2nbfc, national biodiversity future center, palermo, italy abstract mediterranean coastal systems, particularly those characterized by river mouths, have played an important role in the history of human civilization, thanks to unique species assemblages, a high biodiversity, and complex ecological interactions. the overlapping of multiple cross-scale effects, triggered by the ongoing climate change, makes these systems particularly important sites for ecological studies. here, we have investigated the phytoplankton community structure in the terminal stretch of the sele river (tyrrhenian sea) under two different hydrological regimes: i) with river mouth occluded by sediments preventing water exchange with the sea; ii) with river mouth completely opened, which allowed the maximum water mass exchanges between the river and the sea. the strong variations in physical-chemical properties of the water column, the point-distribution of total phytoplankton biomass, and the dominance of cryptophytes and cyanophytes under occluded and opened river mouth, respectively, indicate the presence of strong allogenic pressures acting on the system. our data suggest that the monitoring of phytoplankton communities in the terminal stretch of micro-tidal estuaries, could be a useful tool for studying the extent of climate change underway in mediterranean coastal marine areas. corresponding author: olga mangoni, department of biology, university of naples federico ii, complesso universitario di monte sant’angelo, via cinthia 21, 80126, naples, italy. e-mail: olga.mangoni@unina.it key words: anthropogenic pressure, chemotaxonomy, drought, sele river, tyrrhenian sea. authors’ contributions: all the authors made a substantive intellectual contribution. all the authors have read and approved the final version of the manuscript and agreed to be held accountable for all aspects of the work. conflict of interest: the authors declare no potential conflict of interest. funding: none. availability of data and materials: all data generated or analyzed during this study are included in this published article. acknowledgments: the authors wish to thank the kind collaboration of the natural reserve authority sele-tanagro, fernando guerra for providing and skippering the boat, “la foce dei tramonti camping” for logistic facilities, matilde di somma and erica errico for their support in sampling activities. received: 24 december 2022. accepted: 3 may 2023. publisher’s note: all claims expressed in this article are solely those of the authors and do not necessarily represent those of their affiliated organizations, or those of the publisher, the editors and the reviewers. any product that may be evaluated in this article or claim that may be made by its manufacturer is not guaranteed or endorsed by the publisher. ©copyright: the author(s), 2023 licensee pagepress, italy advances in oceanography and limnology, 2023; 14:11070 doi: 10.4081/aiol.2023.11070 this work is licensed under a creative commons attribution noncommercial 4.0 international license (cc by-nc 4.0). no nco mm er cia l u se on ly f. bolinesi et al.44 seasons depends on the water retention time (lancelot and muylaert, 2011; lichter et al., 2011; newman et al., 2005; pinckney et al., 1999;). many authors state that changes in river water regime from lotic to lentic are accompanied by an increase in allogenic phytoplankton biomass (lampert and sommer, 1997; reynolds, 1984, 1999, 2003; reynolds and descy, 1996; waylett et al., 2013) generating blooms under optimal environmental conditions dominated by species able to grow at intermediate salinity as well as tolerant to large salinity fluctuations in brackish waters of terminal stretches (jackson et al., 1987; muylaert et al., 2009; roubeix and lancelot, 2008). being passively transported by currents, phytoplankton at the source of the river would require a very rapid replication time so that they can generate blooms. this occurs when net specific growth rates of species exceed the residence time of the water (lucas et al., 2009; reynolds, 1984). hence, phytoplankton represents a significant component of primary production only in slow-moving rivers in lowland regions. the responses of the phytoplankton community to variations in water flow rate, nutrient loads, light availability, and temperature are species-specific and highly dependent on the whole environmental context, differing from time to time within each segment of an estuary (artigas et al., 2014). several studies report that diatoms are the dominant phytoplankton group in rivers, with changes in size spectra composition and species associated with changes in nutrient concentration, turbidity, and flushing rate (genkal 1997; reylonds et al., 1994b). for example, schuchardt and schirmer, (1991) and muylaert et al., (2000) report the dominance of diatoms in turbid estuaries throughout the year under high nutrient availability. marshall et al., (2006) report the dominance of diatoms in spring in estuaries where turbidity is low, when light levels are relatively low and stratification does not occur. among other groups, dinoflagellates and cyanobacteria are typically more abundant in summer during periods of low water discharge, long water residence time, minimal flushing rates, and low turbidity levels (lancelot and muylaert, 2011). tavernini et al., (2011) report the dominance of diatoms, chlorophytes, and cyanophytes in the terminal stretch of the po river, with changes in seasonal species assemblages showing an increase of cryptophytes in late july and of cyanophytes at high water discharge rates, from the end of april to mid-june. within the problematic framework described above, we have analyzed the phytoplankton community structure through the detection of diagnostic pigments (ansotegui et al., 2001; mangoni et al., 2017) in the final stretch of the sele river (tyrrhenian sea) under two distinct hydrological regimes: during a severe drought period in july 2017, in which, for the first time in living memory, the mouth of the river was completely occluded by sediments, and in may 2021, in presence of a high water flow rate and a completely open river mouth. our aim was to understand to what extent the strong environmental variations influenced the structure of the phytoplankton community in the terminal stretch of one of the most important river southern italy. materials and methods study area the sele river is the second largest river in southern italy in terms of average water volume after the volturno. it is located within a wide alluvial coastal plain, with a drainage basin of 3235 km2 and a solid flow of 500,000 m3 y-1 (cocco et al., 1989). 64 km long and tributary of the tyrrhenian sea (di paola et al., 2014), it has the highest mean annual flow discharge among the rivers of southern italy (i.e., ~69.4 m3 s-1) (magliulo et al., 2021). it has been an important site for ancient populations, with the first archaeological settlements dating back to the second millennium b.c. (ferrara et al., 2010). the climate of its basis is characterized by prolonged warm and dry summers and wet and mild winters, with mean annual precipitation between 700 to 2000 mm, (average of 1180 mm) showing a marked spatial variation (diodato et al., 2011). over the past 150 years, the sele estuary coastline has been affected by erosion. from 1870 to 1984, the coastline gradually receded, with the highest erosion rates occurring in the main areas around the river mouth. in the last 150 years, the sele river and its catchment area have also been affected by increasing human activities that exerted a significant influence on the evolution of the area. the construction of an artificial drainage channel (canale di bonifacimento) (alberico et al., 2012a, b), with a dense net of artificial drainage channels and the presence of the dam of persano strongly enhanced coastal erosion and altered the river environmental conditions (alberico et al., 2012a, 2012b; arienzo et al., 2020; pappone et al., 2011;). the dam of persano, located 16.2 km from the river mouth, built between 1929 and 1932, created a basin of 1.5 million m3, which affected the downstream sediment deposition causing the consequent retreat of the coast (d’acunzi et al., 2008; freeman et al., 2007; magdaleno et al., 2018). in most recent years, the significant geomorphological alterations of the sele estuary and the presence of pollutants and conspicuous loads of organic matter deriving from anthropogenic activities have been documented, along with a pronounced accumulation of allogenic substances under low flow conditions caused by rainfall deficit (arienzo et al., 2020; benassai et al., 2015; de rosa et al., 2022; montuori et al., 2022;). sampling and environmental parameters water sampling was carried out in the terminal stretch of the sele river, from the mouth up to 3.5 km upstream, under two distinct flow-rate conditions (figure 1a-c). the first sampling was carried out in july 2017 (arienzo et al., 2020) with the river occluded by sediments, a condition that persisted for 1 week, and the second sampling was carried out in may 2021, with the river’s mouth completely opened. stations were located taking into consideration the presence of drainage channels flowing into the river (table 1). at each station, 5 l of water were collected with a niskin bottle and successively subsampled for the analyses of biological and chemical parameters, as described below. for the determination of total phytoplankton biomass (chl a), 50-500 ml of water (depending on the presence of particles in the samples) were drawn from the niskin bottle and filtered onto gf/f whatman glass-fiber filters (25-mm diameter) immediately cryopreserved in liquid nitrogen until the analyses in the laboratory. at each station, water temperature and salinity were determined using idromar xmar212 (2017) and sbe 19 plus (2021) ctd probes interfaced with a gps (garmin map 78s, garmin, olathe, usa). no nco mm er cia l u se on ly phytoplankton community as monitoring tool in the terminal stretch of a micro-tidal estuary facing the tyrrhenian sea 45 total phytoplankton biomass and chemo-functional groups for the determination of total phytoplankton biomass (chl a), 50-500 ml of water (depending on the presence of particles in the samples) were drawn from the niskin and filtered on gf/f whatman filters (25-mm diameter) immediately cryopreserved in liquid nitrogen until the analyses in laboratory (holm-hansen et al., 1965; lorenzen and jeffrey, 1980; yentsch and menzel, 1963). chlorophyll-a and phaeopigments were determined, after extraction with 90% acetone, with a spectrofluorometer (mod.rf– 6000; shimadzu corporation, kyoto, japan), with calibration curves checked daily with fresh chl a standard solutions (sigmaaldrich, st. louis, usa) (holm-hansen et al., 1965). the pheao:chl a ratio was used as a proxy of grazing activity on phytable 1. sampling station with coordinates and bottom depth. station latitude n longitude e depth (m) 1 40.4814990 14.9448934 3.037 2 40.4812454 14.9488489 6.017 3 40.4810787 14.9506658 3.819 4 40.4812777 14.9538731 3.221 5 40.4829086 14.9582883 3.333 6 40.4850757 14.9623601 3.479 7 40.4877350 14.9655997 3.157 8 40.4902284 14.9688533 3.306 9 40.4920247 14.9718688 3.414 10 40.4962250 14.9724679 2.732 11 40.4997458 14.9718300 3.559 figure 1. final stretch of the sele river with sampled stations (a). mouth of the river in july 2017 (b) and may 2021 (c). no nco mm er cia l u se on ly f. bolinesi et al.46 toplankton cells (shuman and lorenzen, 1975). for the determination of the pigmentary spectra, l.5 l of water was filtered on gf/f whatman filters (47-mm dimeter) and the filters were cryopreserved as for chl a. frozen filters were homogenized and resuspended in 100% methanol and analyzed by high-performance liquid chromatography (hplc) (1100 series, hewlett packard, palo alto, usa) in a reverse phase (c8 column 3 μm hyperloop mos) (vidussi et al., 1996). the use of hplc in the study of phytoplankton communities has been largely demonstrated to be a useful tool for the estimation of phytoplankton community composition through the analyses of photosynthetic pigments (jeffery and vesk, 1997; wright et al., 1996). the method is based on the analysis of accessory pigments, in addition to chlorophyll-a (chla) or the modified divinyl-chl-a found in all phytoplankton species, and on the evidence that some of these accessory pigments are taxon-specific (brunet and mangoni, 2010; millie et al., 1997; wright and jeffrey, 2006). this technique allows to detect and identify microscopically overlooked or undetermined ultraphytoplankton species (ansotegui et al., 2003; antajan et al., 2004; garibotti et al., 2003; saggiomo et al., 2023), providing reproducible results. for the determination of chlorophylls and carotenoids, a spectrophotometer with a diode array detector was set at 440 nm, making it possible to determine the absorption spectrum of the 350−750 nm interval for each peak to check the purity of single pigments. the column was calibrated using different pigment standards (chlorophyll a, chlorophyll b, chlorophyll c1+c2, chlorophyll c3, mgdvp, alloxanthin, diatoxanthin, prasinoxanthin, 19′-butanoyloxyfucoxanthin, fucoxanthin, 19′-hexanoyloxyfucoxanthin, peridinin, zeaxanthin, neoxanthin, violaxanthin, bcarotene). standards were provided by the international agency for 14 c determination, vki water quality institute, copenhagen, denmark. quantification was based on the absorbance at 440 nm and the factor response value (peak area/pigment concentration) for each pigment, as described by mantoura and repeta (1997). the contribution of main phytoplankton groups to the total chl a was estimated by chemtax 1.95 software as indicated by latasa (2007), using an iterative process to find the optimal pigment:chl a ratios. the chemotaxonomic groups identified include cyanophytes (cyano), chlorophytes (chloro), prasinophytes (prasino), euglenophytes (eugleno), cryptophytes (crypto), diatoms (diato), pelagophytes (pelago) haptophytes (hapto), dinoflagellates (dino), xantophytes (xanto). inorganic nutrient concentrations for the determination of inorganic nutrient concentrations (n-no3, p-po4), 10 ml water were filtered on 0.2 μm cellulose acetate filters, and stored in hdp vials at -20c. the analyses in the laboratory were conducted using a discrete sampling analyzer (easychem plus, systea, anagni, italy) equipped with a uv-vis spectrophotometer following the procedure described by hansen and grasshoff (1983) adapted to current instrumentation, using a standardized method to detect nitrate (epa ref. national environmental methods index 9171 nitrate via v(iii) reduction), and orthophosphate (iso 15923 ref. international standard organization 15923-1 water quality). statistical analyses the spatial distribution of physical variables (temperature, salinity) was plotted using odv 5.6.3 software. to estimate the pairwise dissimilarity between the two contrasting water flow conditions in a low-dimensional space, a non-metric multidimensional scaling based on a distance matrix computed with the bray-curtis dissimilarity has been performed and represented using a bi-plot in which the correlation coefficients between environmental variable and the nmds scores were presented as vectors from the origin based on algorithm proposed by taguchi and oono (2005). the significance of differences between the two putative conditions of river’s water flow was tested by anosim (clarke, 1993). all statistical analyses were performed using the past 1.95 software. results environmental parameters values of salinity are reported in figure 2. in 2017 the distribution of salinity showed the presence of a strong water stratification, with a net halocline between 1.5-2 m (station 2) and 2-2.5 m (station 9) whose depth increased moving toward the inner part of the river (figure 2a). in general, values ranged between 39.8 and 0.7. at station 1, near the sandbar occluding the river mouth, salinity reached values up to 39 originating a ‘bubble-like’ front probably related to the seawater percolation or episodic waves intrusion from the adjacent sea. the isohalines between 5 and 8, which have been the subject of dispute and debate over time (deaton and greenberg, 1986; khlebovich, 1968, 1969), took place between 1 m (station 2) and 1.8 m depth at station 9, with a slightly deepening moving from the mouth to the inner part of the river. station 1 was the only one with salinity >8 in the entire water column. in 2021, with the river mouth completely opened, the entire sampling area was still characterized by a net water stratification, with a 5-8 salinity boundary cline placed between 1.5 and 2.5 m depth, deepening moving from the mouth (station 1) to the inner part of the river (station 11) (figure 2b). in the entire area, salinity ranged between 36.2 and 0.01. all stations presented two distinct layers: a freshwater one, between 0 and 1.5 m, and a marine one, below 2.5 m. the only exception was station 10, where values reached maximum salinity of 10 on the sea bottom, and the freshwater layer reached its maximum thickness. as far as the temperature is concerned, it showed clear differences between the two sampling periods (figure 3). in july 2017 water temperature exceeded 28.5 °c in the first meter of the water column, generating a marked thermocline at ~1.3 m depth in the overall sampling area (figure 3a). the bottom layer was characterized by the presence of the coldest water reaching the minimum of 24.59 °c at station 2. the surface layer showed a slight increasing trend of temperature moving from the mouth to station 11, with values ranging between 28.5°c and 31.34°c. in may 2021, the water column appeared to be rather homeothermic (figure 3b). in general, values decreased moving from the mouth to the station 11, where temperature reached the minimum of 19.3 °c. a weak hermos-stratification was present in the first 30 cm between stations 1 and 5, where the water retention time reached its maximum. nevertheless, a slightly increased temperature in the bottom water layer was observed in vertical profiles between stations 5 and 9, in accordance with the increase of salinity linked to seawater inflow. the concentration of n-no3 and p-po4 in july 2017 led to no nco mm er cia l u se on ly phytoplankton community as monitoring tool in the terminal stretch of a micro-tidal estuary facing the tyrrhenian sea 47 a mean n:p ratio of 12.12 in the entire area, showing a pointdistribution depending on the influence of drainage channels (arienzo et al., 2020). in particular, n-no3 ranged between 12.85 μmol l-1 (station 5) to 23.80 (station 10) with a mean concentration of 18.2 μmol l-1 ; p-po4 ranged between 0.81 μmol l-1 (station 3) and 8.39 μmol l-1 (station 10) with a mean concentration of 2.84 μmol l-1 (figure 4a). in may 2021, values of n-no3 showed a mean concentration of 15.76 μmol l-1 ranging between 4.57 μmol l-1 (station 2) and 23.51 μmol l-1 (station 9), and displaying a point-shaped distribution, as observed in the first sampling (figure 4b, c). ppo4 were characterized by very low concentrations, with a mean of 0.2 μmol l-1 and values ranging between 0.08 μmol l-1 (station 2) and 0.39 μmol l-1 (station 6). these aspects led to a mean n:p ratio of 83, a value far from what was observed in 2017 and from the redfield one. in particular, the n:p ratio showed values ranging between 148 (station 3) and 30 (station 5), with the highest values in the surface layer. phytoplankton biomass and chemo-functional groups the distribution of total phytoplankton biomass (chl a) in july 2017 showed the presence of eutrophic condition in the figure 2. distribution of salinity along the water column of the entire sampled area in july 2017 (a) and may 2021 (b). y axis: depth (m); x axis: station distance; z colored axis: salinity. no nco mm er cia l u se on ly f. bolinesi et al.48 overall sampling area (figure 5a). values ranged between a minimum of 15.05 µg l-1 at station 9, and a maximum of 95.73 µg l-1 at station 10, highlighting a strong variability in the area, especially near drainage channels. this aspect was also emphasized by the fluctuation of grazing index values (phaeo:chl a), which showed a mean of 0.59 ranging between 0.20 (station 11) and 2.54 (station 9). the overall picture that emerged from chl a concentration indicated the presence of an increasing trend moving from the mouth to the inner part of the river. in may 2021 (figure 5b, c), values of chl a were at least one order of magnitude lower compared to what was reported before, with values usually observed in oligotrophic/mesotrophic systems. the mean value for the entire area was 1.40 µg l-1, with concentrations ranging between 0.55 µg l-1(station 10) and 3.11 µg l-1(station 11). the grazing index ratios were very high at all stations, with a mean of 1.21 and values ranging between 0.62 (station 2) and 1.87 (station 10). as far as the pigment spectra composition (expressed as pigment:chl a ratios) is concerned, we observed high variability in both sampling periods (table s1). in july 2017, fucoxanthin was the dominant pigment, with a mean ratio of 0.67 and values ranging between 0.43 (station 5) and 0.97 (station 1). alloxanfigure 3. distribution of temperature along the water column of the entire sampled area in july 2017 (a) and may 2021 (b). y axis: depth (m); x axis: station distance; z colored axis: temperature (°c). no nco mm er cia l u se on ly phytoplankton community as monitoring tool in the terminal stretch of a micro-tidal estuary facing the tyrrhenian sea 49 thin showed the highest ratio at stations 5 (0.64) and 10 (0.64) and, together with chl c2 and diadinoxanthin was among the most represented ones. chl c3, peridinin, phaeophytin, and 19’hf were the only pigments completely absent in the sampling area. in may 2021, fucoxanthin and mgdvp were the most abundant pigments, with mean ratios of 0.49 and 0.41, respectively. afucoxanthin ranged between 0.36 (station 2, bottom layer) and 0.56 (station 5, bottom layer); mgdvp between 0.29 (station 4) and 0.59 (station 8, bottom layer). the contribution of fucoxanthin and mgdvp to the total pigmentary spectra was higher at the bottom layer than 0 m. neoxanthin was the only pigment to be absent at all stations. peridinin and neoxanthin were completely absent only at surface layer, while other pigments showed weak and nonlinear changes as highlighted by colored bars. the pigmentary spectra composition descripted so far, determined a different chemotaxonomically composition of phytoplankton community in the two sampling periods (figures 6a-c). in july 2017, cryptophytes strongly dominated the community in the entire area, with a mean of 58% and values ranging between 29% (station 2) and 79% (station 11). diatoms and chlorophytes were the second most representative groups, with a mean of 16% and 11% respectively, and higher percentages at station 7. dinoflagellates and haptophytes were completely absent, while xanthophytes were only reported with a weak percentage (2%) at station 11. in may 2021 (figure 7b-c), the community was dominated by cyanophytes and diatoms, with mean percentages of 28.5% and 27.3% respectively. cyanophytes showed a strong gradient at station 4, ranging between 9% (surface layer) and 43% (bottom layer); diatoms ranged between 6% (station 3, bottom layer) and 47% (station 11). chlorophytes and euglenophytes showed similar mean percentage (7%), with the first group that was more represented at the surface layer, and the second one more represented at deeper layer where reaches percentages up to 16% (station 11). haptophytes was the only groups to be absent at surface layer and to be present exclusively at bottom of the station 4. figure 4. box plots with nutrient concentrations of p-po4 and n-no3 in july 2017 (a); surface layer in may 2021 (b); bottom layer in 2021 (c). figure 5. chart bar with chlorophyll a concentration (chl a μg/l) and grazing index (phaeo:chl a) at each station (x axis) in july 2017 (a); surface layer may 2021(b); bottom layer in may 2021(c). no nco mm er cia l u se on ly f. bolinesi et al.50 discussion mediterranean microtidal estuaries have played an important role in human civilization (haidvogl 2018). the present study aims at providing the first insights into changes in phytoplankton community structure in the terminal stretch of the sele river, a microtidal estuary subjected to multiple anthropogenic pressures, as the result of contrasting conditions of the river’s water flow. the reclamation works carried out in the region during the last century have profoundly altered its morphology and degree of interactions between the river, the surrounding lands, and the sea. the extent of these changes is reported today in several geomorphological studies, though with a considerable lack of information on the ecology of the area, especially regarding microalgal communities in the terminal stretch, where phytoplankton plays a pivotal role as primary producers in lentic regime waters. our data represent the first attempt to describe the phytoplankton community structure in one of the most important hydrographic basins of southern italy, under two distinct hydrological phases: with the mouth of the river completely occluded by sediments for the first time in living memory–preventing the water exchange with the sea, and under a high hydrological regime with mouth completely opened. under two different hydrological conditions, the terminal stretch of the sele river was characterized by the coexistence of two distinct water layers. a shallow one, showing typically freshwater salinity whose thickness tends to be higher in the inner part figure 7. non-metric multidimensional scaling based on bray-curtis index. environmental variables are shown in green; the correlation coefficients between each environmental variable and the nmds scores are presented as vectors from the origin. stations of july 2017 are reported as black circles; stations of may 2021 are reported as filled (bottom layer) and empty (surface) orange squares. figure 6. chemotaxonomically composition of phytoplankton community in july 2017 (a); surface layer in may 2021 (b), bottom layer in may 2022 (c). no nco mm er cia l u se on ly phytoplankton community as monitoring tool in the terminal stretch of a micro-tidal estuary facing the tyrrhenian sea 51 of the river, and a deeper one with salinity values that typically exceeded that of the adjacent seawater, especially in deepest ponds and during summer. these thickness changes were accompanied by changes of 5-8 isohaline depth, which was absent near the mouth occluded by sediments and highest at station 10 in the inner part of the river. salinity variation represents a key factor for the biotic component of ecosystems and the presence of brackish organisms is a typical property of estuarine environments. changes in salinity over river mouths depends on the tides, coastal flooding, and season of the year (reynolds, 1984; 2006), so the formation of sandy bar occluding the mouth and preventing the water exchange with the sea, can have significant effects on the functioning and ecology of such a complex system, with multi-scale effects still unknown. since dramatic rainfall-decreasing trends were reported for the southern italian tyrrhenian coast, causing droughts during summer (polemio and casarano, 2004, 2008; romano et al., 2022), the results presented here could be the first sign of a new condition towards which this system will likely tend more frequently in the near future. contrarily to the salinity, the thermal profile of the water column showed marked differences between the two sampling periods in accordance with the season. in july 2017, a net thermocline was present at ~ 1.3 m depth, with water temperature exceeding 28.5 °c in the first meter of the water column reaching colder values on the bottom (minim 24.59 °c, at station 2). in spring 2021 the water column was rather homeothermic, with lower values in the inner part of the river and relatively warmer water on the bottom linked to the seawater intrusion. a weak stratification was observed in the first 30 cm between stations 1 and 5, and despite the fact current meters data are not available, we hypothesize that this warmer layer reflects the hydrological dynamics of the area, where the increase in water retention time favors surface waters warming. it must be noted that stations were sampled ~10 minutes apart from each other, moving from the mouth to station 11, so sampling activities should not have influenced these data. many authors indicate that any potential alteration to river flows would increase the temporal and spatial variability of estuarine fronts, water stratification, and mixing, with potentially negative impacts associated with eutrophication (reynolds, 1984; 2006). the impact of nutrient enrichment on rivers is often season specific and complicated by their dynamic nature (newman et al., 2005). differences in nutrient concentration observed in our study, with extremely large spatial and temporal variations in n:p ratios, suggest how complex and articulated are the dynamics of nutrient supply in this area, where the presence of drainage channels (in addition to the combined effects of water retention times) can strongly influence the chemical properties of the water column (albanese et al., 2007; arienzo et al., 2020; montuori et al., 2022). nitrates did not show drastic changes between the two sampling periods, as instead observed with the phosphate load, which led to higher n:p ratios in 2021, with values up to 148 at station 3, compared to those in 2017. the first biological effect of nutrient enrichment in pelagic environments is the growth of phytoplankton species, whose blooms extent and community structure strongly depend on local conditions, for example, the extent to which the nutrient load is retained by the receiving water (newman et al., 2005; reynolds, 2006). this is the reason why rivers are considered highly selective environments, where flow rate and turbidity, along with temperature, are the most critical factors limiting the development and structuring the phytoplankton community composition (rojo et al., 1994; reynolds, 1994, tavernini et al., 2011). the higher the flow rate and turbidity, the lower the amount of phytoplankton can be expected: where the water retention time is longer than the generation time of the phytoplankton, large blooms can occur (ibeling et al.,1998; lampert and sommer, 1997; newman et al., 2005). nevertheless, the extent to which nutrient loads trigger phytoplankton blooms in micro-and macrotidal estuaries is different, with small nutrient enrichment producing larger effects in the microtidal ones (warwick et al., 2018). in general, allogenic factors (i.e. inoculum of phytoplankton species transported downstream, temperature, light regime, discharge rate, turbidity) are considered the main drivers of phytoplankton succession in lowland rivers (reynolds, 1994, 2006), with autogenic ones becoming more relevant with increasing eutrophication levels (del giorgio et al., 1991). the distribution of total phytoplankton biomass (chl a) in july 2017 points to a speckled distribution with an increasing trend of concentration moving toward the inner part of the river, where values are typical of eutrophic waters reaching the maximum of 97.73 μg l-1. the station with lowest concentrations of chl a (15.0 μg l-1) showed the highest phaeo:chl a ratio (2.7), indicating the presence of large amounts of degraded biomass, most likely because of the adjacent drainage channel transport. on the contrary, in may 2021, values of chl a were at least one order of magnitude lower, with a mean of 1.40 μg l-1, appearing more homogeneously distributed in the entire sampling area and being characterized by very high phaeo:chl a ratios, the latter supporting the hypothesis that much of the phytoplanktonic biomass was photosynthetically inactive. the overall picture emerging from nmds analyses clearly shows the differences between the two samplings periods, with data from july 2017 (black points) clustering in the left side of the plot, and those from may 2021 (orange squares) in the right part. temperature, p-po4, and the n:p ratios are among the main environmental drivers shaping the phytoplankton community, with high chl a concentration and cryptophytes characterizing the first sampling period, and high n:p ratios characterizing the second one. in may 2021 (orange squares), euglenophytes, cyanophytes, and xantophytes were the dominant groups, especially at the surface layer (filled orange squares). the correlation between salinity and depth highlights the role of seawater inflow in shaping the structure of the phytoplankton communities, with dinoflagellates, haptophytes, and pelagophytes dominating in the deepest and saltiest layers, and high loads of n-no3 that were instead associated to the shallow layer, because of inputs from drainage channels. a high and positive r value (0.9777) in the anosim analysis (p=0.0001) indicates the strong and significant dissimilarity between the phytoplankton communities’ composition in the two sampling periods (table 2). altogether our data table 2. results of the anosim (analyses of similarities) test carried out to ascertain differences between the two sampling periods, representing contrasting water flow conditions. anosim test – bray-curtis permutation n 9999 mean rank within 132 mean rank between 334.6 r 0.9777 p (same) 0.0001 no nco mm er cia l u se on ly f. bolinesi et al.52 agree with the results reported by tavernini et al., (2011), who, in the po river, observed the highest concentration of cryptophytes biovolume in late july, concomitantly with an higher contribution of cyanophytes between april and june: these patterns provide a further confirmation that the terminal stretch of the sele river is exposed to strong allogenic pressures. conclusions in addition to what discussed so far, we pinpoint here that the phytoplankton community represents a promising and effective monitoring tool for the study of estuarine environments, in which cross-scale interactions produce large fluctuations in physical, chemical, and biological properties. nevertheless, the large variability of the phytoplankton community structure both in space and time, and the extremely ample variations in n:p ratios underline how much allogenic pressures can influence communities’ functional traits in the terminal stretch of the sele river and, supposedly, in other estuarine areas, with similar combinations of natural and anthropogenic pressure. the presence in both study periods of a saltier layer near the bottom highlights the role played by the seawater inflow in the ecology of the area, characterized by a strong summer thermoaline stratification, which exacerbates the difference between superficial and bottom water layers. although our results are not sufficient to mechanistically explain the functioning of such a complex ecosystem, the ample array of interactions among local climate, hydrodynamical and ecological variables that emerged from our study suggest that the sele river estuary will most likely be increasingly exposed to the multiple stressors caused by the ongoing climate change and the increase of human uses of land and rivers. the presence of other estuarine environments in the mediterranean sea and the peculiar characteristics of the terminal stretches of rivers altogether suggest that more effort should be paid to investigate and monitor these delicate and vulnerable ecosystems. this appears to be urgently needed in order to identify new management measures of freshwater inputs in the mediterranean sea, one of the marine regions worldwide most prone to the consequences of climate change. references albanese s, de vivo b, lima a, cicchella d, 2007. geochemical background and baseline values of toxic elements in stream sediments of campania region (italy). j. geochem. explor. 93, 21–34. alberico i, amato v, aucelli ppc, et al., 2012b. historical shoreline change of the sele plain (southern italy): the 1870-2009 time window. journal of coastal research, 28:1638-47. alberico i, amato v, aucelli cpp, et al., 2012a. historical and recent changes of the sele river coastal plain (southern italy): natural variations and human pressures. rend. fis. acc. lincei, 23:3–12. ansotegui a, sarobe a, trigueros jm, et al., 2003. size distribution of algal pigments and phytoplankton assemblages in a coastal—estuarine environment: contribution of small eukaryotic algae. journal of plankton research. 25:341–55. ansotegui a, trigueros jm, orive e, 2001. the use of pigment signatures to assess phytoplankton assemblage structure in estuarine waters. estuarine, coastal and shelf science. 52:689-703. antajan e, chrétiennot-dinet m,-j, leblanc c, et al., 2004. 19′hexanoyloxyfucoxanthin may not be the appropriate pigment to trace occurrence and fate of phaeocystis: the case of p. globosa in belgian coastal waters. journal of sea research. 52:165–77. arienzo m, bolinesi f, aiello g, et al., 2020. the environmental assessment of an estuarine transitional environment, southern italy. journal of marine science and engineering. 8:628. artigas ml, llebot c, ross on, et al., 2014. understanding the spatio-temporal variability of phytoplankton biomass distribution in a microtidal mediterranean estuary. deep sea research part ii: topical studies in oceanography. 101:180-92. benassai g, di paola g, aucelli ppc, 2015. coastal risk assessment of a micro-tidal littoral plain in response to sea level rise. ocean & coastal management. 104:22-35. brunet c, mangoni o, 2010. metodologie di campionamento e di studio del plancton marino. in g. socal, i. buttino, m. cabrini, o. mangoni, a. penna, & c. totti (eds.), manuali e linee guida (vol. 56, pp. 343–349). ispra. caloiero t, caroletti gn, coscarelli r, 2021. imerg-based meteorological drought analysis over italy. climate. 9:65. clarke kr, 1993. non-parametric multivariate analysis of changes in community structure. australian journal of ecology. 18:117-143 cocco e, de magistris ma, de pippo t, et al., 1989. coastal dynamics along the shores of campania and lucania (southern italy). in: proceedings of the sixth symposium on coastal and ocean management, charleston, sc, usa, 3, pp. 2794–2807. d’acunzi g, de pippo t, donadio c, et al., 2008. studio dell’evoluzione della linea di costa della piana del sele (campania) mediante l’uso della cartografia numerica. stud. costieri. 14:55–67. davies br, walker kf, 1986. the ecology of river systems. monographiae biologicae, 60. 793 pp. deaton le, greenberg mj, 1986. there is no horohalinicum. estuaries. 9:20–30. del giorgio pa, vinocur al, lombardo rg, tell hg, 1991. progressive changes in the structure and dynamics of the phytoplankton community along a pollution gradient in a lowland river—a multivariate approach. hydrobiologia. 224:129–54 de rosa e, montuori p, triassi m, et al., 2022. occurrence and distribution of persistent organic pollutants (pops) from sele river, southern italy: analysis of polychlorinated biphenyls and organochlorine pesticides in a water–sediment system. toxics. 10:662. di paola g, aucelli ppc, benassai g, german r, 2014. coastal vulnerability to wave storms of sele littoral plain (southern italy). natural hazards. 71:1795–819. diodato n, fagnano m, alberico i et al., 2011. mapping soil erodibility from composed data set in sele river basin, italy. nat hazards. 58:445–57. dugdale r, wilkerson f, parker ae, et al., 2012. river flow and ammonium discharge determine spring phytoplankton no nco mm er cia l u se on ly phytoplankton community as monitoring tool in the terminal stretch of a micro-tidal estuary facing the tyrrhenian sea 53 blooms in an urbanized estuary. estuarine, coastal and shelf science. 115:187-99. ferrara b, 2010. heraion alla foce del sele: l’uso del territorio tra scienza e archeologia. scienze naturali ed archeologia: il paesaggio antico: interazione uomo /ambiente ed eventi catastrofici. atti convegno internazionale, p. 123-126. freeman mc, pringle cm, jackson cr, 2007. hydrologic connectivity and the contribution of stream headwaters to ecological integrity at regional scales. j. am. water res. assoc. 43:5–14. garibotti ia, vernet m, kozlowski wa, ferrario me, 2003. composition and biomass of phytoplankton assemblages in coastal antarctic waters: a comparison of chemotaxonomic and microscopic analyses. marine ecology progress series. 247, 27–42. gasith a, and resh vh, 1999. streams in mediterranean climate regions: abiotic influences and biotic responses to predictable seasonal events. annual review of ecology and systematics. 30:51-81. genkal si, 1997. problems in identifying centric diatoms for monitoring the water quality of large rivers. in: j. prygiel, b.a. whitton, j. bukowska (eds), use of algae for monitoring rivers iii. p. 182-187. douai, france, agence de l’eau artois-picardie. haidvogl g, 2018. historic milestones of human river uses and ecological impacts. riverine ecosystem management. 2018:19-39. haralambidou k, sylaios g, tsihrintzis va, 2010. salt-wedge propagation in a mediterranean micro-tidal river mouth. estuarine, coastal and shelf science. 90:174-184. hearn cj, and robson bj, 2001. inter-annual variability of bottom hypoxia in shallow mediterranean estuaries. estuar. coast. shelf sci. 52:643–57. holm-hansen o, lorenzen cj, holmes rw, strickland jdh, 1965. fluorometric determination of chlorophyll. ices journal of marine science. 30:3-15. ibelings b, admiraal w, bijkerk r, et al., 1998. monitoring of algae in dutch rivers: does it meet its goals? journal of applied phycology. 10:171-81. jackson rh, williams pjl, joint ir, 1987. freshwater phytoplankton in the low salinity region of the river tamar estuary. estuarine, coastal and shelf science. 25:299–311. jeffrey sw, vesk m, 1997. introduction to marine phytoplankton and their pigment signatures. khlebovich vv, 1968. some peculiar features of the hydrochemical regime and the fauna of mesohaline waters. mar. biol. 2:47-9. khlebovich vv, 1969. aspects of animal evolution related to critical salinity and the internal state. mar. biol. 2:338-45. kennish m, 2002. environmental threats and environmental future of estuaries. environmental conservation. 29:78-107. kennish m, 2021. drivers of change in estuarine and coastal marine environments: an overview. open journal of ecology. 11:224-39. lampert w, sommer u, 1997. limnoecology: the ecology of lakes and streams. oxford university press, new york, usa. 336 pp. lancelot c, muylaert k, 2011. 7.02 trends in estuarine phytoplankton ecology. available from: https://www.researchgate. net/profile/arvind-singh-21/post/did-you-find-any-relationship-between-the-water-quality-and-distribution-of-phytoplankton-in-that-estuarine-waters/attachment/59d64e4379197 b80779a7a31/as%3a492057329442816%40149432701745 3/download/235306.pdf legendre p, legendre l, 1998. numerical ecology, 2nd english ed. elsevier, the netherlands. 853 pp. lichter m, klein m, zviely d, 2011. dynamic morphology of small south-eastern mediterranean river mouths: a conceptual model. earth surf. process. landforms. 36:547-62. lorenzen cj, jeffrey sw, 1980 determination of chlorophyll in sea water. unesco tech. pap. mar.sci. 35:1-20. lucas lv, thompson jk, brown lr, 2009. why are diverse relationships observed between phytoplankton biomass and transport time? limnology and oceanography. 54:381–90. magdaleno f, donadio c, kondolf gm, 2018. 30 year response to damming of a mediterranean river in california, usa. phys. geogr., 39:197–215. magliulo p, cusano a, russo f, 2021. land-use changes in the sele river basin landscape (southern italy) between 1960 and 2012: comparisons and implications for soil erosion assessment. geographies. 1:315–32. mangoni o, saggiomo v, bolinesi f, et al. 2017. phytoplankton blooms during austral summer in the ross sea, antarctica: driving factors and trophic implications. plos one. 12:e0176033. mantoura rfc, repeta dj 1997. calibration methods for hplc. in: jeffrey sw, mantoura rfc, wright sw (eds) phytoplankton pigments in oceanography: guidelines to modern methods. unesco, paris, france. pp. 407. marshall hg, lacouture rv, buchanan c, johnson jm, 2006. phytoplankton assemblages associated with water quality and salinity regions in chesapeake bay, usa. estuarine, coastal and shelf science. 69:10–8. millie df, schofield om, kirkpatrick gj, et al., 1997. detection of harmful algal blooms using photopigments and absorption signatures: a case study of the florida red tide dinoflagellate, gymnodinium breve. limnology and oceanography. 42: 1240–51. montuori p, de rosa e, di duca f, et al., 2022. occurrence, distribution, and risk assessment of organophosphorus pesticides in the aquatic environment of the sele river estuary, southern italy. toxics. 10:377. muylaert k, sabbe k, vyverman w, 2000. spatial and temporal dynamics of phytoplankton communities in a freshwater tidal estuary (schelde, belgium). estuarine, coastal and shelf science. 50:673–87. muylaert k, sabbe k, vyverman w, 2009. changes in phytoplankton diversity and community composition along the salinity gradient of the schelde estuary (belgium/ the netherlands). estuarine, coastal and shelf science. 82:335–40. newman j, anderson nj, bennion h, et al., 2005. eutrophication in rivers: an ecological perspective. technical report. adas, 37 pp. pappone g, alberico i, amato v, et al., 2011. recent evolution and the present-day conditions of the campanian coastal plains (south italy): the case history of the sele river coastal plain. wit trans. ecol. environ. 149:15–27. perillo me, 1995. chapter 2. definitions and geomorphologic classifications of estuaries. g.m.e. perillo (ed), developments in sedimentology. 53:17-47. no nco mm er cia l u se on ly f. bolinesi et al.54 pinckney jl, paerl hw, harrington mb, 1999. responses of the phytoplankton community growth rate to nutrient pulses in variable estuarine environments. journal of phycology. 35:1455-63. polemio m, casarano d, 2004. rainfall and drought in southern italy (1821-2001). iahs-aish publication. 286:217–27. polemio m, casarano d, 2008. climate change, drought and groundwater availability in southern italy. in: w. dragoni and b.s. sukhija (eds), climate change and groundwater. the geological society special publications. 288:39-51. reynolds cs, 1984. phytoplankton periodicity: the interactions of form, function and environmental variability. freshwater biology. 14:111–42. reynolds cs, 1994a. the long, the short and the stalled: on the attributes of phytoplankton selected by physical mixing in lakes and rivers. hydrobiologia. 289:9–21. reynolds cs, 1994b. the role of fluid motion in the dynamics of phytoplankton in lakes and rivers. in: p.s. giller, a.g. hildrew, d. raffaelli (eds), ecology of aquatic organisms: scale, pattern, process. blackwell scientific publications, oxford, uk, p.141–87. reynolds cs, 1999. modelling phytoplankton dynamics and its application to lake management. hydrobiologia. 395:123–31. reynolds cs, 2003. planktic community assembly in flowing water and the ecosystem health of rivers. ecological modelling. 160:191–203. reynolds cs, 2006. the ecology of phytoplankton. cambridge university press, cambridge, uk 535 pp. reynolds cs, descy jp and padisák j, 1994. are phytoplankton dynamics in rivers so different from those in shallow lakes? hydrobiologia. 289:1-7. reynolds cs, descy jp, 1996. the production, biomass and structure of phytoplankton in large rivers. archiv für hydrobiologie, supplementband large rivers. 10:161–87. romano e, petrangeli ab, salerno f, guyennon n, 2022. do recent meteorological drought events in central italy result from long-term trend or increasing variability? international journal of climatology. 42:4111–28. rojo c, cobelas ma, arauzo m, 1994. an elementary, structural analysis of river phytoplankton. hydrobiologia. 289:43–55. roubeix v, lancelot c, 2008. effect of salinity on growth, cell size and silicification of an euryhaline freshwater diatom cyclotella meneghiniana kutz. transitional waters bulletin. 1:31–8. saggiomo m, bolinesi f, brunet c, et al., 2023. a chemtaxderived phytoplankton community structure during 12-year observations in the gulf of naples (lter-mc). marine ecology. 2023:e12745. schuchardt b, schirmer m, 1991. phytoplankton maxima in the tidal freshwater reaches of two coastal plain estuaries. estuarine, coastal and shelf science. 32:187–206. shuman f, lorenzen r, carl j, 1975. quantitative degradation of chlorophyll by a marine herbivore. limnology and oceanography. 20:580-6. taguchi y-h, oono y, 2005. relational patterns of gene expression via non-metric multidimensional scaling analysis. bioinformatics 21:730-40. tavernini s, pierobon e, viaroli p, 2011. physical factors and dissolved reactive silica affect phytoplankton community structure and dynamics in a lowland eutrophic river (po river, italy). hydrobiologia. 669:213–25. ustin ssm, hestir e, khanna s, et al., 2014. developing the capacity to monitor climate change impacts in mediterranean estuaries. evolutionary ecology research. 16:529–50. van der tuin h, 1991. guidelines on the study of seawater intrusion into rivers. international hydrological programme, unesco, 138 pp. valle-levinson a, 2010. definition and classification of estuaries. available from: https://assets.cambridge.org/97805218/ 99673/excerpt/9780521899673_excerpt.pdf verri g, pinardi n, oddo p, et al., 2018. river runoff influences on the central mediterranean overturning circulation. clim. dyn. 50:1675–703. vidussi f, claustre h, bustillos-guzman j, et al., 1996. determination of chlorophylls and carotenoids of marine phytoplankton: separation of chlorophyll a from divinyl chlorophyll a and zeaxanthin from lutein. journal plankton research. 18:2377-82. warwick rm, tweedley rj, potter ic, 2018. microtidal estuaries warrant special management measures that recognise their critical vulnerability to pollution and climate change, marine pollution bulletin. 135:41-46. waylett aj, hutchins mg, johnson ac, et al., 2013. physicochemical factors alone cannot simulate phytoplankton behaviour in a lowland river. journal of hydrology. 497:223–33. weisse t, 2008. limnoecology: the ecology of lakes and streams. journal of plankton research. 30:489-90. whitton ba, 1975. river ecology. vol. 2. university of california press, berkley, usa. 725 pp. yentsch cs, menzel dw, 1963. a method for the determination of phytoplankton chlorophyll and phaeophytine by fluorescence. deep sea res.10:221-31. wright sw, jeffrey sw, 2006. pigment markers for phytoplankton production. marine organic matter: biomarkers, isotopes and dna. 2005;71–104. wright sw, thomas dp, marchant hj, et al., 1996. analysis of phytoplankton of the australian sector of the southern ocean: comparisons of microscopy and size frequency data with interpretations of pigment hplc data using the “chemtax” matrix factorisation program. marine ecology progress series. 144:285–98. online supplementary material: table s1. pigments: chl a ratios at each station in july 2017 (a); surface layer may 2021(b); bottom layer in may 2021(c). no nco mm er cia l u se on ly layout 1 introduction since its first application, the use of environmental dna (edna) isolated from water samples to detect the presence of taxa has been considered a promising method to improve aquatic biomonitoring (ficetola et al., 2008; goldberg et al., 2015; lawson handley, 2015). environmental dna is the genetic material present in different environmental matrices such as sediment, water, and air, and belonging to the organisms inhabiting the surveyed area; it includes dna released in the environment (intra or extracellular) and dna taken directly from living cells (pawlowski et al., 2018). this edna can be extracted from the environmental matrices and used to assess community biodiversity through the amplification of a short dna region used as a “barcode” (hebert et al., 2003; taberlet et al., 2012; ward et al., 2009) and sequenced with high-throughput sequencing techniques (i.e., metabarcoding, kuntke et al., 2020; schenekar et al., 2020). the use of edna metabarcoding can significantly improve biodiversity monitoring surveys through the early detection of exotic and potentially invasive species and the tracking of elusive endangered species (deiner et al., 2018; pawlowski et al., 2018; taberlet et al., 2018). a range of organisms is used worldwide as indicators (biological quality elements (bqes)) to monitor the quality status of aquatic ecosystems, namely phytoplankton, phytobenthos, aquatic plants, macroinvertebrates, and fish (european environment agency, 2018). the sampling method officially recommended for the biodiversity assessment of river and lake fish by the water framework directive (en14011, 2003) are electrofishing and gill netting, which are quite expensive methods and require a large and qualified staff to be performed. moreover, several sampling practices (e.g., seines or trawling) can severely damage the habitat and in large lakes can heavily affect fish abundance (irvine et al., 2019; njiru et al., 2018). since 2000, the european union has been actively engaged in the protection and enhancement of aquatic ecosystems: freshwater biomonitoring promoted by the implementation of the eu water framework directive (wfd) (european commission, 2000) allows for the evaluation and improvement of their quality status. however, classical biomonitoring practices require good taxonomic expertise and the probability of detecting species that represent <1% of the total abundance is article alpine freshwater fish biodiversity assessment: an inter-calibration test for metabarcoding method set up giulia riccioni1*, isabelle domaizon2, andrea gandolfi1, massimo pindo1, adriano boscaini1, marine vautier2, hans rund3, peter hufnagl4, stefanie dobrovolny4, valentin vasselon5, jonas bylemans1,6, cuong q. tang7, nico salmaso1, josef wanzenböck3 1research and innovation centre, fondazione edmund mach, s. michele all'adige, italy; 2national institute of research for agriculture, alimentation and environment, university savoie mont blanc, pole r&d ecla, thonon-les-bains, france; 3university of innsbruck, research department for limnology, mondsee, austria; 4austrian agency for health and food safety gmbh, wien, austria; 5french office for biodiversity, pole r&d ecla, thonon-les-bains, france; 6university of lausanne, department of ecology and evolution, biophore, lausanne, switzerland; 7naturemetrics ltd., egham, uk abstract the analysis of environmental dna (edna) by high throughput sequencing (hts) is proving to be a promising tool for freshwater fish biodiversity assessment in europe within the water framework directive (wfd, 2000/60/ec), especially for large rivers and lakes where current fish monitoring techniques have known shortcomings. these new biomonitoring methods based on edna show several advantages compared to classical morphological methods. the sampling procedures are easier and cheaper and edna metabarcoding is non-invasive and very sensitive, allowing for the detection of traces of dna. however, edna metabarcoding methods need careful standardization to make the results of different surveys comparable. the aim of the eu project eco-alpswater is to test and validate molecular biodiversity monitoring tools for aquatic ecosystems (i.e., edna metabarcoding) to improve the traditional wfd monitoring approaches in alpine waterbodies. to this end, an inter-calibration test was performed using fish mock community samples containing either tissue-extracted dna, edna collected from aquaculture tanks and edna samples collected from lake bourget (france). samples were analysed using a dna metabarcoding approach, relying on the amplification and hts of a 12s rdna marker, in two separate laboratories, to evaluate if different laboratory and bioinformatic protocols can provide a reliable and comparable description of the fish communities in both mock and natural samples. our results highlight good replicability of the molecular laboratory protocols for hts and good amplification success of selected primers, providing essential information concerning the taxonomic resolution of the 12s mitochondrial marker in describing the alpine fish communities. interestingly, different concentrations of species dna in the mock samples were well represented by the relative dna reads abundance. these tests confirm the reproducibility of edna metabarcoding analyses for the biomonitoring of freshwater fish inhabiting alpine and peri-alpine lakes and rivers. no nco mm er cia l u se on ly g. riccioni et al.12 very low (paller, 1995), thus providing partial estimates for entire communities (deiner et al., 2017). compared to electrofishing and gill netting procedures, water sampling procedures for edna analyses proved to be potentially cheaper and easier, non-invasive, and suited for surveys in extremely difficult sites. reduced operational costs could allow for regular sampling during the year, providing time-series data and a systematic monitoring of fish biodiversity (and, in general, of community diversity) in different seasons and extreme events (extremely dry seasons or floods). conversely, traditional sampling can usually be performed once or twice a year. standardization of all the protocols, from the sampling activities to the taxonomic assignment of dna sequences (dickie et al., 2018), is paramount to allow for the comparability among edna metabarcoding studies for the ecological assessment of habitats or ecosystems (goldberg et al., 2015) and to improve the sensitivity of metabarcoding assays. the use of experimental controls and mock communities allows to exclude unspecific signals and verify the recovery of species signals and quantitative representation of species in the samples, as well as to determine bioinformatic filtering steps, and threshold levels. using standardized methods allows for a better interpretation of ecosystem response to pressures (mock and kirkham, 2012; morales and holben, 2011) however, each study has its specificity and often requires a customization of the metabarcoding protocols that often need further validation, especially when universal polymerase chain reaction (pcr) primers are used to explore fish community diversity (bylemans et al., 2018; thalinger et al., 2021a). pilot studies, comprising inter-calibration tests, are an invaluable tool to evaluate if different approaches can lead to similar and comparable results (zinger et al., 2019). it is for these reasons that, in this study, we performed an inter-calibration exercise involving two different laboratories to evaluate the reproducibility and possible limits of an edna metabarcoding procedure to describe the fish community diversity in alpine lakes and rivers. no previous studies were performed to estimate alpine fish biodiversity by using edna metabarcoding, and a pilot test was highly recommended. mock samples made of tissue dna pools as well as aquaria water samples and lake water samples were included in the test to compare the species detection performance and replicability of the method. this test has been conceived within the eco-alpswater network, a project funded by the european union, whose ultimate goal is to evaluate and validate emerging technologies based on edna metabarcoding for the biodiversity assessment of freshwater ecosystems in the alpine region (https://www.alpine-space.eu/projects/eco-alpswater). materials set-up of mock and environmental samples all the samples were prepared at inrae (institut national de recherche pour l’agriculture, l’alimentation et l’environnement) as a blind test for the laboratories involved in the inter-calibration exercise. a negative control sample (pure water filtered following the same protocol used for the other samples) was included in the test and analysed following the same procedures (dna extraction, library preparation, and sequencing) used for the other samples. three mock samples including an increasing number of species pooled (6, 9, and 14, hereafter m6, m9, and m14) and different dna proportions (table s1) were set up using genomic dna extracts from fin clips. we used one individual per fish species to set up these three mock samples. the fish species used to assemble these mock samples are species commonly found in lakes and rivers of the alpine and perialpine regions. a fourth mock sample was prepared by collecting 100 ml of water from each of ten fish tanks containing one single species each (for a total of seven species, namely salmo carpio l., oncorhynchus mykiss w., barbus caninus b., perca fluviatilis l., lepomis gibbosus l., carassius carassius l., and tinca tinca l.). for each tank, we collected the same volume of water to obtain a total volume of 1 l to simulate a real environmental sampling procedure. this sample was collected on the 3rd of december, 2019, and extracted on the 9th of december, 2019. moreover, a third set of three environmental samples of 6 x 2 l each were collected from three different areas in lake bourget (all subsurface samples, i.e., 10-20 cm; figure 1). three sampling points were sampled from each lake bank to simulate a transect and 2 l of water were collected from each site (figure 1). after careful mixing of the samples, 1 l of water collected from the fish tank sample and lake bourget samples were filtered by using 0.45 µm sterivextm capsule filters (merck millipore, burlington, usa); the filtration cartridges were filled with spygen (spygen, le bourget du lac cedex, france) preservation buffer and stored at room temperature until the dna extraction step. these lake samples were collected on the 16th of october, 2019, stored at room temperature, and extracted on the 21st and 22nd of october, 2019. dna extraction and library preparation for fin clips dna extractions, we used the nucleospin® dna rapidlyse kit from macherey-nagel. edna extraction was performed using the nucleospin® soil kit (macherey-nagel, allentown, usa) following the protocol described in pont et al. (2018) and adapted to sterivextm capsule filters. twenty µl of genomic dna no nco mm er cia l u se on ly edna metabarcoding test for fish biodiversity 13 extract were delivered to both fondazione edmund mach, italy (lab_a) and naturemetrics, uk (lab_b) sequencing platforms for further laboratory processing. for both laboratories, a hypervariable region of 12s rrna was amplified via a two-step pcr process (supplementary information). the two laboratories performed all pcrs in the presence of both a negative and a positive control (i.e., a mock community with a known composition of fish species). amplification success at each step was determined by gel electrophoresis. all pcrs replicates per sample were pooled and purified using cleanngs beads (cleanna, waddinxveen, netherlands) by lab_a and magbind totalpure ngs (omega biotek, norcross, usa) magnetic beads with a ratio of 0.8:1 (beads: dna) by lab_b, to remove primer dimers. the sequences, saved in fastq formatted files, were deposited to the european nucleotide archive (ena) with study accession number prjeb49223. bioinformatic analyses bioinformatic analyses were performed by lab_a using the obitools3 software (boyer et al., 2016). a reference database was created simulating a pcr amplification in silico by using the ecopcr program: the whole vertebrate embl database (june 2020) and the mifishu primers were used, allowing for a maximum of three mismatches with the published sequences, and blocking the last two nucleotides in the primer sequence. for fastq miseq sequences analyses the alignpairedend script was used to perform a micro-assembly of pairedend reads and sequences with illumina fastq quality scores <30 across the head, tail, or total length of the sequence were discarded. the ngsfilter script was used to assign the reads to each sample through barcode identification and, after a dereplication step, only sequences longer than 80 nucleotides and a count ≥10 were retained for further analyses. the obiclean script was used to detect the potential pcr errors, selecting only sequences with the ‘head’ status and abundance higher than 0.05%. the taxonomic assignment was performed by using the ecotag script and the reference database, considering a 97% of similarity. the taxonomic assignment was further inspected by using blastn (zhang et al., 2000) algorithm optimized for very similar sequences (megablast) on the nucleotide collection (nr/nt) that includes all genbank + embl + ddbj + pdb sequences when uncertainties in the identification emerged. rarefaction analysis was performed on lake bourget sequences to evaluate if the sequencing effort allowed to reach a plateau using the vegan package (oksanen, 2016) and the rarecurve function in r environment (r core team, 2020). to evaluate the intra-individual variability of the 12s rdna copies, both salmo trutta miseq sequences assigned at the species level and those assigned only at the genus level and obtained from the mock tissue sample (dna extracted from the fin clip of a single individual), were further evaluated by using megax (kumar et al., 2018). these sequences were aligned separately by using the muscle software (edgar, 2004) with default parameters, and the mean distance among sequences was computed by using the kimura 2-parameter model (k2p, kimura, 1980) and 1000 bootstrap. at lab_b samples were demultiplexed based on the combination of the i5 and i7 index tags. paired-end reads for each sample were merged with usearch (edgar, 2010), with a minimum overlap of 20% of the total read length. forward and reverse primers were trimmed from the merged sequences using cutadapt (martin, 2011), and retained if the trimmed length was between 140 and 200 bp. these sequences were quality filtered with usearch to retain only those with an expected error rate per base of 0.05 or below, and dereplicated by sample, retaining singletons. unique reads from all samples were denoised in a single analysis with unoise (edgar and flyvbjerg, 2015), requiring retained zotu`s (zero-radius otu’s) to have a minimum abundance of 8 in at least one sample. a taxon-by-sample table was generated by mapping all dereplicated reads for each sample to the zotu representative sequences with usearch, at an identity threshold of 97%. zotu’s were identified via blastn (zhang et al., 2000) searches of the representative sequences against the whole nt database and a local curated database of 12s fish sequences. identifications were based figure 1. sampling sites map of lake bourget. three 6 x 2 l samples were collected in three different areas of the lake and 1 l of water was filtered to evaluate the effectiveness of the edna metabarcoding survey in describing the fish community diversity (boub = low lake bourget, boum = medium lake bourget, bouh = high lake bourget). no nco mm er cia l u se on ly g. riccioni et al.14 on the highest available percentage identity at 98–100%, with an e-score of 1e-20 and a hit length of at least 80% of the query sequence. in cases where multiple reference sequences matched equally to the query sequence then a more conservative higher taxonomic classification was considered. only sequences with speciesor genus-level identifications were included in the final results. when a species was represented by multiple zotus, the one with the highest percentage match to that species was taken as the representative. typically, the other sequences having the same occurrence pattern and the lower sequence similarity can be attributed to pcr or sequencing errors. data analysis analysis of covariance (ancova) was used to compare two regression lines by testing the effect of a categorical factor (the two laboratories involved in the high throughput sequencing (hts) analyses) on the dependent variable (fraction of taxonomic annotated sequences) while controlling for the effect of the continuous covariable (fraction of dna of single species in mock tissue samples; crawley, 2005). the results allowed to test differences in the regression slopes (interaction effect) and intercepts (main effects) in the two regression models (kéry, 2010). moreover, a regression analysis was performed by using the ggpubr function in r to compare the number of otus obtained by the two laboratories for the lake bourget and the data (fish number and biomass) collected from the traditional survey performed in lake bourget in 2018 (table s11, figure s2). statistical analyses were computed using r 4.03 (r core team, 2020). for these analyses, only the sequences identified at the species or genus level by both lab_a and lab_b were considered. results lab_a analyses produced 14,600,000 sequences and, after the quality control step, an average of 250,000 sequences per sample were obtained, excluding the negative control sample. after obitools3 analyses, 1,500,000 sequences were assigned, and 20 species, 9 genera, and 3 families respectively were identified. the blank sample produced less than 5000 sequences that did not find any match with vertebrate sequences present in the embl database, with a similarity threshold of 97%. lab_b analyses produced an average of 320,000 sequences per sample (excluding the template negative control). after lab_b quality control steps, an average of 270,000 sequences per sample were achieved. a total of 1,900,000 sequences after lab_b bioinformatic analyses were successfully assigned, identifying 22 species of fish. the blank sample failed to amplify and yielded less than 500 sequences. tissue dna mock samples the identification success of the tissue dna mock sample revealed good performance of both the lab_a and lab_b analyses, with few inconsistencies in the assignment of otus to coregonus, esox, cyprinus and salmo genera (table 1). more in detail, coregonus lavaretus l. sequences were assigned only at the genus level by obitools3, because the 12s rdna fragment is not diagnostic for species within this genus, and the ecotag script assigns a sequence to the most recent common ancestor when a sequence shows the same percentage of similarity with different species, as shown by a blast similarity search (table s4): five different species were indeed identified with 100% of identity. moreover, some of these sequences were assigned only at the family level, suggesting the presence of different genera among the reference sequences showing a level of similarity higher than 97% (as proved by a blastn similarity search, table s5). the same sequence was assigned to 14 different species belonging to the coregonus genus and one species belonging to the stenodus genus, with a percentage of identity between 98 and 99%. the same inconsistencies are found for cyprinus, esox, salmo and silurus (tabs s6, s7, s8 and s9). some squalius cephalus sequences were assigned to leuciscus sp. because of a limited resolution power of the 12s fragment, as revealed by a blastn search, revealing a sequence similarity between 98 and 99% to both squalius cephalus and leuciscus leuciscus of sequences identified as leuciscus sp. lab_b assigned a few coregonus sequences to coregonus maraena b., which is a species distributed in the baltic sea basin and was not included in the mock sample (table 1). furthermore, a wrong salvelinus species was identified (salvelinus fontinalis m.) instead of the expected salvelinus alpinus l. notably, three species present in very low proportions (table 1) in the m14 mock sample were not detected by both the lab_a and lab_b procedures, namely gobio gobio l., esox lucius l. and lota lota l. the comparison between the dna proportions included in the mock samples for each species and the dna sequences proportions retrieved using hts for these species showed a good correlation (for both laboratories, figure 2). this was also confirmed for the mock assemblage with the most complex species composition (m14). in the three mock assemblages, the individual slopes were always significant (p<0.001) and ranged between 0.8 and 1.1. moreover, both slopes and intercepts did not show significant differences (p>0.05). the mean genetic k2p distance computed among the salmo trutta l. sequences assigned at the species level was <0.01, whereas the k2p distance calculated among the sequences identified only as salmo genus was 0.02. no nco mm er cia l u se on ly edna metabarcoding test for fish biodiversity 15 mock communities in the fish tanks the analyses performed on the dna collected from fish tanks (table 2, figure 3) showed that some species can be identified by lab_a analyses only at the genus level because still missing from the embl database (i.e., barbus caninus and salmo carpio) and, for some species, the fragment of the 12s marker amplified by mifish primers is not informative (see oncorhynchus mykiss, table s10). moretable 1. species identification of mock samples made of dna extracted from fish fin clips. for m6, m9 and m14 samples 6, 9 and 14 species were pooled respectively with different proportions of dna (see table s1). in bold are highlighted the assignment errors. lab_a analysis lab_b analysis inrae mock samples inrae mock samples species present species identified m6 m9 m14 m6 m9 m14 abramis brama (m6, m9, m14) abramis brama 22926 10147 13607 18950 15261 14656 ameiurus melas (m14) ameiurus melas 0 0 0 0 0 648 coregonus lavaretus (m6, m9, m14) coregonus maraena 0 0 0 14566 11089 22707 coregonus sp. 13315 7380 21001 0 0 0 coregoninae 2495 1263 3828 0 0 0 cyprinus carpio (m9, m14) cyprinus carpio 0 0 0 0 19272 4010 cyprinidae 5452 7880 4378 0 0 0 esox lucius (m9, m14) esox lucius 0 18327 0 0 40949 0 esox sp. 0 2153 0 0 0 0 not present leuciscus sp. 1130 449 675 0 0 0 perca fluviatilis (m6, m9, m14) perca fluviatilis 64823 29446 95258 69639 57379 94871 rutilus rutilus (m14) rutilus rutilus 84 37 230 0 0 585 salmo trutta (m6, m9, m14) salmo trutta 53479 31158 30666 60915 55216 40149 salmo sp. 18439 9415 11116 0 0 0 salmo labrax 19 0 15 0 0 0 salvelinus alpinus (m6, m9, m14) salvelinus fontinalis 0 0 0 22446 19829 36366 salvelinus sp. 32463 17043 44745 0 0 0 salmoninae 11735 6261 10338 0 0 0 silurus glanis (m9, m14) silurus glanis 0 0 0 0 4231 798 silurus sp. 0 56 14 0 0 0 squalius cephalus (m6, m9, m14) squalius cephalus 48656 20671 28778 47551 35022 35302 tinca tinca (m14) tinca tinca 0 0 399 0 0 991 table 2. taxonomic assignment results of the metabarcoding analyses of dna collected from fish tanks. in bold are highlighted the assignment errors. lab_a analysis lab_b analysis water tank sample water tank sample species present species identified barbus caninus barbus ciscaucasicus 240 0 carassius carassius cyprinus carpio 0 1589 cyprinidae 891 0 engraulis ringens 70 0 lepomis gibbosus lepomis gibbosus 6394 17451 oncorhynchus mykiss 67055 54680 oncorhynchus sp. 13361 0 oncorhynchus mykiss oncorhynchus clarkii henshawi 30 0 oncorhynchus nerka 24 0 perca fluviatilis perca fluviatilis 17376 47373 salmo trutta 144596 107995 salmo sp. 34515 0 salmo carpio salmo labrax 79 0 salmoninae 3555 0 tinca tinca tinca tinca 6059 17533 no nco mm er cia l u se on ly g. riccioni et al.16 over, lab_a analyses detected the presence of engraulis ringens, a species often used to produce a fish meal for aquaculture. all the species present in the fish tank sample were detected by the two laboratories, except barbus caninus, which was not identified by lab_b analyses even at the genus level, and carassius carassius, whose 12s sequence is not taxonomically informative (table s6). the venn diagram highlighted that 4 out of 7 species were correctly identified by the two laboratories (figure 3). the three species missed by the molecular analysis were still absent in the embl database or not informative. lake bourget samples the metabarcoding analyses performed on the environmental samples collected in lake bourget showed a good agreement between lab_a and lab_b (figure 4). only for the rare species represented with a low number of sequences some discrepancies emerged (e.g., barbatula barfigure 2. relationship between the fraction of assigned high throughput sequencing (hts) and the fractions of fish mock dna pooled in different proportions. each point represents 1 different species, as represented in tab 1. the graphs refer to mock assemblages with mix of (a) 6, (b) 9 and (c) 11 species; in (c), three species with a very low amount of dna were excluded from the analysis; see tab 1 for details about the species included in the three graphs. in the legend, lab_a and lab_b refer to the two sequencing facilities. in the three mock assemblages (graphs a-c), all the regression lines were highly significant (p<0.001), whereas both slopes (around 1) and intercepts did not show significant differences. (d) relationships between the proportions of assigned sequences detected in lab_a and lab_b from the analysis of fish mock dna (r2=0.92, p<0.001). no nco mm er cia l u se on ly edna metabarcoding test for fish biodiversity 17 batula l., gasterosteus aculeatus l., pseudorasbora sp., oncorhynchus mykiss). the taxonomic identification allowed for the detection of 23 species or genera from the sequences processed by lab_a, and 20 species from the sequences processed by lab_b. in total, 13 fish species were detected by both laboratories (without considering the species assigned only at the genus level; figure 4). as for the other taxa, some species were assigned by lab_a only at the genus level, and the analyses assigned the sequences to different species within a genus (e.g., salmo trutta). when comparing the hts results obtained from the three sampling stations by both lab_a and lab_b with the current records of species detected in the whole lake bourget by using the traditional sampling approach (mostly electrofishing and gill netting from 1995 to 2018), 17 and 19 species respectively out of 35 were correctly detected (table 3, table 4). when considering only the most recent survey, performed in 2018 with traditional gear, 13 out of 15 species were identified by the edna metabarcoding approach (table s11). the regression analyses performed on otus numbers, and traditional survey data (fish number and biomass, table s11) showed significant r values for all the comparisons (figure s2). the rarefaction curves constructed by using the sequences obtained from lake bourget samples showed that the sequencing effort could describe the diversity present in the three samples as all the three curves reached the plateau (figure s1). by limiting the comparison only to the assigned sequences identified at the species or genus level by both lab_a and lab_b, the results confirmed the high degree of comparability of hts data (figure 5). the spearman correlations between the fractions of otus abundances figure 3. venn diagram of the edna metabarcoding taxonomic identification obtained from the samples collected from the fish tanks at fem fishery facility. lab_a = fondazione edmund mach, lab_b =nature metrics. figure 4. venn diagram of the edna metabarcoding taxonomic identification of the water samples collected in lake bourget. lab_a = edmund mach foundation, lab_b = nature metrics. figure 5. relationship between the fractions of assigned sequences obtained from the high throughput sequencing (hts) analyses of the three environmental samples collected in lake bourget determined in lab_a and lab_b. each point represents 1 different species. the comparison is limited to the common species found in the two sets of analyses. boub = low lake bourget, boum = medium lake bourget, bouh = high lake bourget. no nco mm er cia l u se on ly g. riccioni et al.18 computed on the total of the respective samples obtained in lab_a and lab_b for the three sets of samples ranged between 0.84 and 0.85 (0.001<p<0.03). discussion in this study, an inter-calibration test has been applied to evaluate the effectiveness of an edna metabarcoding protocol in estimating fish community biodiversity in lakes and rivers of the alpine region. our results confirmed the need for prior evaluation of the laboratory and bioinformatic protocols by using mock pooled samples of known species composition to validate the bioinformatic procedures, to verify if the protocols used can identify all the species present (and what is the detection threshold) and how reliable the results obtained are. these analyses also allowed to determine the limits of the method and its reproducibility among different laboratories. previous surveys have been performed to test the ability to reconstruct the species composition of fish communities using mock samples or aquarium samples (doi et al., 2019; miya et al., 2015; morey et al., 2020; thalinger et al., 2021b; turanov and rutenko, 2020). however, this is the first attempt to reconstruct freshwater fish communities inhabiting the alpine area using both mock tissue samples and tanks samples. table 3. taxonomic assignment of sequences obtained from the edna collected in three different sites in lake bourget and analysed by lab_a and lab_b metabarcoding protocols. in bold are highlighted the assignment errors. lab_a analysis lab_b analysis bourget samples bourget samples species identified boub bouh boum species identified boub bouh boum abramis brama 692 0 1093 abramis brama 2953 316 4362 ameiurus melas 534 546 0 barbatula barbatula 0 176 0 barbatula barbatula 612 2302 0 barbus barbus 256 0 175 cottus gobio 1334 0 0 coregonus maraena 3375 1202 8709 coregonus sp. 4085 2370 18572 coregoninae 203 125 2068 cyprinus carpio 70289 307 2349 cyprinidae 35812 137 1908 esox lucius 5843 5005 12986 esox lucius 10326 9827 22440 esox sp. 364 287 1097 esox flaviae 0 0 12 gasterosteus aculeatus 0 244 0 gobio gobio 500 0 0 gymnocephalus cernua 1692 292 0 gymnocephalus cernua 503 419 0 gymnocephalus sp. 18 0 0 lepomis gibbosus 31 795 4123 lepomis gibbosus 890 4199 5216 leuciscus leuciscus 6050 892 0 leuciscus sp. 3400 638 0 leuciscus sp. 901 0 0 oncorhynchus mykiss 0 0 953 oncorhynchus sp. 0 0 26 perca fluviatilis 68625 96369 78289 perca fluviatilis 142050 182363 158519 pseudorasbora sp. 437 0 0 rutilus rutilus 17284 25574 43938 rutilus rutilus 26461 62385 65468 salmo trutta 0 0 0 salmo trutta 78 143 120 salmo labrax 0 0 896 salmo sp. 0 0 44 scardinius erythrophthalmus 0 0 414 scardinius erythrophthalmus 794 473 0 silurus glanis 462 0 4812 silurus sp. 41 0 0 squalius cephalus 639 429 0 squalius cephalus 2697 0 0 tinca tinca 24294 31811 25105 tinca tinca 24713 49709 39313 no nco mm er cia l u se on ly edna metabarcoding test for fish biodiversity 19 consistency of high throughput sequencing with the composition of mock and environmental fish assemblages the pooled dna analysis extracted from fish fin clips clearly showed that the mifish primers targeting the 12s rdna cannot distinguish some species within the coregonus, salmo and esox genus. the amplified fragment is highly conserved in some taxa showing little variation among species/genera, with the result that the taxonomic identification of sequences remains limited to the genus rank. for cyprinus carpio, the blast search highlighted the presence of 12s rdna sequences belonging to different species and genera with a percentage of similarity higher than 97%, which allows an assignment only at the family level. moreover, some sequences belonging to one individual (e.g., salmo trutta in the mock tissue dna sample) were assigned to different species or only at the genus level. this finding suggested the presence of some inter-copies variation of 12s rdna marker within a single individual, as revealed by the k2p distance values, or some remnant pcr/sequencing errors, although a very conservative bioinformatic procedure was used. the analysis performed by lab_b assigned these identical sequences at the species level but identified, in some table 4. comparison between the fish species recorded using traditional sampling in lake bourget (between 1995 and 2018) and edna metabarcoding assessment performed by lab_a and lab_b. for the species not identified by the metabarcoding approach, we have checked the presence of 12s rdna mifish fragments in embl database: 0 = missing, 1 = present. traditional sampling recordsedna metabarcoding lab_aedna metabarcoding lab_bpresence of reference sequence species note taxonomic assignment taxonomic assignment 12s mifish fragment 1 abramis brama abramis brama abramis brama 2 alburnoides bipunctatus 0 3 alburnus alburnus 1 4 ameiurus melas ameiurus melas 5 barbatula barbatula barbatula barbatula barbatula barbatula 6 barbus barbus barbus barbus 7 blicca bjoerkna 1 8 coregonus lavaretus coregonus sp. coregonus maraena 9 cottus gobio cottus gobio 10 cyprinus carpio cyprinus carpio 11 esox lucius esox lucius esox lucius 12 gasterosteus aculeatus gasterosteus aculeatus 13 gobio gobio gobio gobio 14 gymnocephalus cernua gymnocephalus cernua gymnocephalus cernua 15 lampetra planeri 0 16 lepomis gibbosus lepomis gibbosus lepomis gibbosus 17 leuciscus leuciscus leuciscus sp. leuciscus leuciscus 18 lota lota 1 19 oncorhynchus mykiss oncorhynchus mykiss 20 perca fluviatilis perca fluviatilis perca fluviatilis 21 phoxinus phoxinus 1 22 pseudorasbora parva pseudorasbora sp. 23 rhodeus amarus 1 24 rutilus rutilus rutilus rutilus rutilus rutilus 25 salaria fluviatilis 0 26 salmo trutta salmo sp. salmo trutta 27 salvelinus alpinus 1 28 salvelinus fontinalis only once in 1995 and uncertain 1 29 sander lucioperca 0 30 scardinius erythrophthalmus scardinius erythrophthalmus scardinius erythrophthalmus 31 silurus glanis silurus sp. silurus glanis 32 squalius cephalus squalius cephalus squalius cephalus 33 telestes souffia 0 34 thymallus thymallus 1 35 tinca tinca tinca tinca tinca tinca no nco mm er cia l u se on ly g. riccioni et al.20 cases, a wrong species (e.g., coregonus maraena and salvelinus fontinalis). in general, the two laboratories were able to describe the fish community composition at least at the genus level. only three species, namely gobio gobio, esox lucius and lota lota, which were present in very low proportions of dna (<0.01%, table s1), escaped the identification. the fish tank experiments were consistent with the above results, highlighting a good performance of the metabarcoding analyses with some limits due to the absence of reference sequences in the database or the lack of resolution of the fragment of 12s rdna used for the taxonomic identification. also, in this analysis, the presence of intra-individual 12s rdna copy variation emerged for salmo trutta and was suggested for oncorhynchus mykiss as a very low number of sequences were assigned to two different species (oncorhynchus clarkii henshawi and oncorhynchus nerka) and an important number of sequences were assigned at the genus level. in fish tanks, several individuals per species were present and, for this reason, the computation of the intraindividual 12s rdna copy variation was not feasible. the identification of a low number of sequences misassigned for salmo and oncorhynchus suggested that these copy variants could be present in low frequency in the mitochondrial genome of these species. interestingly, lab_a detected a species commonly used as fish meal, namely engraulis ringens, which was probably present only in traces as very few sequences were detected. this outcome confirms the high sensitivity of this method, which can identify traces of dna in a pool of more abundant genetic material, as also revealed by the results obtained from the tissue dna mock sample. the analyses of “real” environmental samples collected in lake bourget confirmed the results of mock samples showing that both lab_a and lab_b approaches provided a comparable description of the fish community in the lake. some discrepancies emerged for a few species that showed a low number of sequences and was probably represented by a very low amount of dna. our test on the tissue dna mock sample showed a threshold level of detection as low as 0.05% (table 1 rutilus rutilus). the taxonomic assignment of salmo sequences highlighted that lab_a detected only salmo labrax or salmo sequences, whereas lab_b identified salmo trutta in the same sample. salmo labrax was never recorded in lake bourget, and is present only in the danube area and eurasia rivers draining to the black sea. by inspecting these sequences, it appears that salmo trutta and salmo labrax sequences differ only for one nucleotide change out of a total of 170 nucleotides of the 12s rdna fragment. this confirms that special attention should be given when considering the taxonomic assignment of sequences belonging to these species and further inspection (also using a blast search) is advisable. notably, the three 6 x 2 l water samples collected in the lake were able to detect 17 and 19 species (lab_a and lab_b, respectively) out of 35 species recorded from 1995 to 2018 through the classical collection and identification methods, and the sequencing effort used revealed to be able to identify all the species present in the samples. this outcome stresses the high potential of this technique for the description of fish biodiversity in lakes and rivers, despite some limits highlighted by this intercalibration test. issues implicit in the taxonomic resolution of the genetic markers should be carefully evaluated and, in this regard, a further taxonomic assignment step using a customised database, including only the target species already recorded in the habitat under study, can help improve the assignment of sequences at the species level (as shown by lab_b results). a few species were not identified because they were not present in the reference databases (table 4). a more comprehensive database for the taxonomic assignment is recommended to detect allochthonous species that would not be included in the customised database of the selected species. moreover, a known limit of the edna approach (and barcoding approach in the broader sense) is the lack of information in public databases that can prevent the identification of some species which can be undetected. a further obvious control of the reliability of the species identified by using the edna approach is the plausibility of the taxonomic composition of the fish community described using a priori information of the ecosystem under study. this information can be obtained for lakes and rivers of the alpine region from the systematic assessment performed by using the classical sampling gears as in the case of lake bourget, described in this work. quantitative assessment of fish communities besides testing the efficiency in detecting single species, the experiments carried out with the tissue dna mock samples allowed us to verify that the potential pcr and sequencing errors do not influence the final proportion of dna sequences assigned for each species. in this regard, the experiments were based on controlled conditions and analyses carried out on predetermined fractions of pure dna extracted from single specimens, therefore removing all the biases that usually are present when “blind” comparisons, without previous knowledge of communities, are carried out in the evaluation of laboratory performances. the hts in the lab_a and lab_b identified the sequences associated with the respective fish species with almost the same proportions as those used to artificially create the three mock samples. further, the two laboratories showed comparable no nco mm er cia l u se on ly edna metabarcoding test for fish biodiversity 21 performances, indicating a good reproducibility of methods. as for this last aspect, and limiting the analyses to the only common species found in the two laboratories, a good reproducibility was obtained by comparing the fractions of the sequence abundances of the species identified taxonomically in the lab_a and lab_b. the above results highlight and further demonstrate the sensitivity of the hts methods in quantifying the number of specific gene markers in heterogeneous genomic dna extracts, although several issues intrinsic to the procedure only allow for a relative estimation of fish biomasses. these include the sample collection and the nature and stochastic distribution of biological tissues and materials released by fish (jo et al., 2019; sassoubre et al., 2016). many environmental parameters, like mixing/dispersion and the time spent by fish in the area, as well as biotic parameters (skin drop off rates and dna decay rates), are indeed important factors to relate edna concentrations to actual fish biomass and abundance. previous studies reported a temperature-dependent degradation of edna (eichmiller et al., 2016; lance et al., 2017; strickler et al., 2015; tsuji et al., 2017), an accelerated edna degradation due to higher water temperatures and higher fish biomass (jo et al. 2019) and a correlation with physiology and behaviour of different species groups (thalinger et al. 2021b). conclusions the edna hts approach to monitor fish biodiversity in alpine lakes and rivers proved to be an efficient tool to complement the current biomonitoring surveys, and the correct implementation of the metabarcoding method and interpretation of the results obtained is crucial to avoid wrong conclusions. this work demonstrated a good comparability of the two laboratories participating in the inter-calibration test results. in particular, the differences in the quantitative estimation of the relative abundances of otus in three fish mock assemblages were insignificant. on the other side, a few relevant differences were apparent in the discrimination and taxonomic annotation at the species or genus level of selected taxa. possibly, one of the most serious issues in the application of hts methods is still represented by the incomplete coverage of species in the nucleotide databases. this has important implications when hts methods are used in the biomonitoring of aquatic environments or for the early detection of non-indigenous species. these drawbacks should be taken into consideration when interpreting edna data. in the case of biomonitoring, it is urgent to complete the coverage of reference databases, at least on a regional basis, including not only native species, but also those recognised as potential non-indigenous species. corresponding author: giulia riccioni. e-mail: giulia.riccioni0276@gmail.com. authors’ contributions: all the authors made a substantive intellectual contribution, performed part of the experiments. all the authors have read and approved the final version of the manuscript and agreed to be accountable for all aspects of the work. conflict of interest: the authors declare no potential conflict of interest. availability of data and materials: all data generated or analyzed during this study are included in this published article. funding: this work has been carried out with the support and cofinancing of the european regional development fund (erdf) through the interreg alpine space program, project eco-alpswater. jb is currently supported by funding from the swiss national science foundation (snf grant number: 31003a_182265). key words: fish biodiversity; edna; alpine space; eco-alpswater; metabarcoding. received: 6 august 2021. accepted: 5 february 2022. publisher’s note: all claims expressed in this article are solely those of the authors and do not necessarily represent those of their affiliated organizations, or those of the publisher, the editors and the reviewers. any product that may be evaluated in this article or claim that may be made by its manufacturer is not guaranteed or endorsed by the publisher. ©copyright: the author(s), 2022 licensee pagepress, italy advances in oceanography and limnology, 2022; 13:10017 doi: 10.4081/aiol.2022.10017 this work is licensed under a creative commons attribution noncommercial 4.0 international license (cc by-nc 4.0). references boyer f, mercier c, bonin a, le bras y, taberlet p, coissac e, 2016. obitools: a unix-inspired software package for dna metabarcoding. mol. ecol. resour. 16:176–182. bylemans j, gleeson dm, hardy cm, furlan e 2018. toward an ecoregion scale evaluation of edna metabarcoding primers: a case study for the freshwater fish biodiversity of the murray–darling basin (australia). ecol. evol. 8:8697–8712. crawley mj, 2005. statistics: an introduction using r. john wiley & sons inc., hoboken, usa: 354 pp. deiner k, bik hm, mächler e, seymour m, lacoursière-roussel a, altermatt f, et al., 2017. environmental dna metabarcoding: transforming how we survey animal and plant communities. mol. ecol. 26:5872–5895. deiner k, lopez j, bourne s, holman l, seymour m, grey ek, et al, 2018. optimising the detection of marine taxonomic richness using environmental dna metabarcoding: the effects of filter material, pore size and extraction method. metabarcoding and metagenomics 2:e28963 no nco mm er cia l u se on ly g. riccioni et al.22 dickie ia, boyer s, buckley hl, duncan, rp, gardner, pp, hogg, id, et al, 2018. towards robust and repeatable sampling methods in edna-based studies. mol. ecol. resour. 18:940–952. doi h, fukaya k, oka s, sato k, kondoh m, miya m, 2019. evaluation of detection probabilities at the water-filtering and initial pcr steps in environmental dna metabarcoding using a multispecies site occupancy model. sci. rep. 9:1–8. edgar rc, 2004. muscle: multiple sequence alignment with high accuracy and high throughput. nucleic acids res. 32:1792–1797. edgar, rc, 2010. search and clustering orders of magnitude faster than blast. bioinformatics 26:2460–2461. edgar rc, flyvbjerg h, 2015. error filtering, pair assembly and error correction for next-generation sequencing reads. bioinformatics 31:3476–3482. eichmiller jj, best se, sorensen pw, 2016. effects of temperature and trophic state on degradation of environmental dna in lake water. environ. sci. technol. 50:1859–1867. en14011, c.e.n., 2003. water analysis-fishing with electricity for wadable and non-wadable rivers. eur. comm. stand. european commission, 2000. water framework directive 2000/60/ec. off. j. eur. communities l 269:1–15. european environment agency, 2018. european waters assessment of status and pressures 2018, eea report no 7/2018. evans nt, li y, renshaw ma, olds bp, deiner k, turner cr, et al, 2017. fish community assessment with edna metabarcoding: effects of sampling design and bioinformatic filtering. can. j. fish. aquat. sci. 74:1362–1374. ficetola, gff, miaud c, pompanon f, taberlet p, 2008. species detection using environmental dna from water samples. biol. lett. 4:423–425. goldberg cs, strickler km, pilliod, ds, 2015. moving environmental dna methods from concept to practice for monitoring aquatic macro-organisms. biol. conserv. 183:1–3. hänfling, b, handley ll, read ds, hahn c, li j, nichols p, et al., 2016. environmental dna metabarcoding of lake fish communities reflects long-term data from established survey methods. mol. ecol. 25:3101–3119. hebert pdn, cywinska a, ball sl, dewaard jr, 2003. biological identifications through dna barcodes. proc. r. soc. london. ser. b biol. sci. 270:313–321. irvine k, etiegni ca, weyl olf, 2019. prognosis for long-term sustainable fisheries in the african great lakes. fish. manag. ecol. 26:413–425. jo t, murakami h, yamamoto s, masuda r, minamoto t, 2019. effect of water temperature and fish biomass on environmental dna shedding, degradation, and size distribution. ecol. evol. 9:1135–1146. kéry m, 2010. introduction to winbugs for ecologists: bayesian approach to regression, anova, mixed models and related analyses. elsevier science publishing co. inc, amsterdam, netherlands: 320 pp. kimura m, 1980. a simple method for estimating evolutionary rates of base substitutions through comparative studies of nucleotide sequences. j. mol. evol. 16:111–120. kumar s, stecher g, li m, knyaz c, tamura k, 2018. mega x: molecular evolutionary genetics analysis across computing platforms. mol. biol. evol. 35:1547. kuntke f, de jonge n, hesselsøe m, nielsen jl, 2020. stream water quality assessment by metabarcoding of invertebrates. ecol. indic. 111:105982. lance rf, klymus, ke, richter ca, guan x, farrington hl, carr mr, et al, 2017. experimental observations on the decay of environmental dna from bighead and silver carps. manag. biol. invasions 8:343. lawson handley l, 2015. how will the ‘molecular revolution’ contribute to biological recording? biol. j. linn. soc. 115:750–766. martin m, 2011. cutadapt removes adapter sequences from highthroughput sequencing reads. embnet.journal 17:10-12. miya m, sato y, fukunaga t, sado t, poulsen jy, sato k, et al, 2015. mifish, a set of universal pcr primers for metabarcoding environmental dna from fishes: detection of more than 230 subtropical marine species. r. soc. open sci. 2:150088. mock t, kirkham a, 2012. what can we learn from genomics approaches in marine ecology? from sequences to eco-systems biology! mar. ecol. 33:131–148. morales se, holben we, 2011. empirical testing of 16s pcr primer pairs reveals variance in target specificity and efficacy not suggested by in silico analysis. handb. mol. microb. ecol. i metagenomics complement. approaches 135–141. morey kc, bartley tj, hanner rh, 2020. validating environmental dna metabarcoding for marine fishes in diverse ecosystems using a public aquarium. environ. dna 2:330–342. njiru j, van der knaap m, kundu r, nyamweya c, 2018. lake victoria fisheries: outlook and management. lakes reserv. sci. policy manag. sustain. use 23:152–162. oksanen j, 2016. vegan: an introduction to ordination. paller mh, 1995. relationships among a number of fish species sampled, reach length surveyed, and sampling effort in south carolina coastal plain streams. north am. j. fish. manag. 15:110–120. pawlowski j, kelly-quinn m, altermatt f, apothéloz-perretgentil l, beja p, boggero a, et al., 2018. the future of biotic indices in the ecogenomic era: integrating (e) dna metabarcoding in biological assessment of aquatic ecosystems. sci. total environ. 637:1295–1310. pont, d., rocle, m., valentini, a. et al. environmental dna reveals quantitative patterns of fish biodiversity in large rivers despite its downstream transportation. sci rep 8, 10361 (2018). https://doi.org/10.1038/s41598-018-28424-8 r core team, 2020. r: a language and environment for statistical computing. sassoubre lm, yamahara km, gardner ld, block ba, boehm ab, 2016. quantification of environmental dna (edna) shedding and decay rates for three marine fish. environ. sci. technol. 50:10456–10464. sato h, sogo y, doi h, yamanaka h, 2017. usefulness and limitations of sample pooling for environmental dna metabarcoding of freshwater fish communities. sci. rep. 7:1–12. schenekar t, schletterer m, lecaudey la, weiss, sj, 2020. reference databases, primer choice, and assay sensitivity for environmental metabarcoding: lessons learnt from a reevaluation of an edna fish assessment in the volga headwaters. river res. appl. 36:1004–1013. no nco mm er cia l u se on ly edna metabarcoding test for fish biodiversity 23 shaw jla, clarke lj, wedderburn sd, barnes tc, weyrich ls, cooper a, 2016. comparison of environmental dna metabarcoding and conventional fish survey methods in a river system. biol. conserv. 197:131–138. strickler km, fremier ak, goldberg cs, 2015. quantifying effects of uv-b, temperature, and ph on edna degradation in aquatic microcosms. biol. conserv. 183:85–92. taberlet p, bonin a, coissac e, zinger l, 2018. environmental dna: for biodiversity research and monitoring. oxford university press, oxford, uk: 268 pp. taberlet p, coissac e, hajibabaei m, rieseberg lh, 2012. environmental dna. mol. ecol. 21:1789–1793. thalinger b, deiner k, harper lr, rees hc, blackman rc, sint d, et al., 2021a. a validation scale to determine the readiness of environmental dna assays for routine species monitoring. environ. dna 3:823-836. thalinger b, rieder a, teuffenbach a, pütz y, schwerte t, wanzenböck j, et al., 2021b. the effect of activity, energy use, and species identity on environmental dna shedding of freshwater fish. front. ecol. evol. 9:1–13. tsuji s, ushio m, sakurai s, minamoto t, yamanaka h, 2017. water temperature-dependent degradation of environmental dna and its relation to bacterial abundance. plos one 12: e0176608. turanov sv, rutenko oa, 2020. the composition of northeast pacific fishes in a fish tank examined by edna metabarcoding. biorxiv 2020.12.21.423745. ward rd, hanner r, hebert pdn, 2009. the campaign to dna barcode all fishes, fish-bol. j. fish biol. 74:329–56. zhang z, schwartz s, wagner l, miller w, 2000. a greedy algorithm for aligning dna sequences. j. comput. biol. 7:203–214. zinger l, bonin a, alsos ig, bálint m, bik h, boyer f, et al., 2019. dna metabarcoding—need for robust experimental designs to draw sound ecological conclusions. mol. ecol. 28:1857–1862. no nco mm er cia l u se on ly layout 1 article dissolved fluxes of nutrients and carbon at the sediment-water interface in the adriatic sea: review of early data and methods from the italian national research council (cnr) federico spagnoli1,2, mariangela ravaioli3 1istituto per le risorse biologiche e le biotecnologie marine, ancona (irbim-cnr); 2università di camerino, school of science and technology, geology division, camerino (mc); 3istituto di scienze marine, bologna (ismar-cnr), italy abstract we report the measurement of the fluxes of dissolved compounds at the sediment-water interface (dfswis), obtained by analyzing data collected in the italian exclusive economic zone (eez) of the northern and central adriatic sea from 1982. we also describe the methods, honed by italy’s consiglio nazionale delle ricerche (cnr), employed to measure and calculate dfswis, including benthic chambers and landers, sediment-water interface micro-profiling, on-board incubation, and pore water modeling. data analysis demonstrated that in situ measurements are the most reliable approaches, but that on-board collected data also supply interesting results. the dfswi data allowed to divide the adriatic seafloor into areas with similar dfswi types and intensities and to define the role of sediments as sources or sinks of chemicals such as nutrients, dissolved inorganic carbon (dic), and metals. dfswis mainly depend on dissolved and solid river inputs, marine organic matter (mom) production, and sediment reworking. they show a seasonal temperature-related trend, which in summer induces increased chemical reaction kinetics and microbial activity. dfswis decline from the po river mouths southward, along the holocene mud wedge. north of the po river delta, the dfswis are weaker and confined to the areas in front of the major river mouths, due to poor organic matter (om) inputs and strong reworking of bottom sediments. the area south of the po delta cusp is characterized by strong dfswis, due to the high deposition of solid inputs from the po river, strong primary production, and protection from the western adriatic current and bora storms; in summer, higher temperatures, and calm hydrodynamic conditions generate near-bottom hypoxic to anoxic environments in this area. in northern and central adriatic offshore areas, negative dic and phosphate fluxes are due to poorly reactive om reaching this area and to the mainly carbonate composition and oxic environment of local bottom sediments. dfswi data analysis highlighted the important contribution of sediments to marine carbon and nutrient cycles. this is particularly important for carbon, which plays a major role in seawater acidification and global climate change. the study also provides average dfswi data for each diagenetic area, which allow calculating the carbon and nutrient budgets in the adriatic sea. corresponding author: federico spagnoli, istituto per le risorse biologiche e le biotecnologie marine, largo fiera della pesca, ancona (irbim-cnr), italy. e-mail: federico.spagnoli@cnr.it key words: benthic fluxes, benthic chambers, pore water modeling, adriatic sea, sediment biogeochemistry, early diagenesis. authors’ contributions: all the authors made a substantive intellectual contribution. all the authors have read and approved the final version of the manuscript and agreed to be held accountable for all aspects of the work. conflict of interest: the authors declare no potential conflict of interest. funding: funding are indicated in the acknowledgements. availability of data and materials: all data generated or analyzed during this study are included in this published article. acknowledgements: we acknowledge several valuable contributions. marco marcaccio, maria cristina bergamini and franca frascari participated in data collection and processing. paola giordani, doug hammond and will berelson were the first to produce data about dfswis in the adriatic sea and introduced the methods for their measurement and calculation. rocco de marco helped with some figures, gabriele marozzi provided some photographs, gabriella rovatti and stefano miserocchi worked in the research groups led by paola giordani. technicians enver lipparini, giuseppe baldelli, angelo magagnoli and maurizio mengoli participated in the cruises and in instrument setup. we are grateful to the students who contributed to data collection and processing. finally, silvia giuliani made a valuable final revision of the paper. we acknowledge funding from the following main projects: “research and experimentation for the adriatic sea” prisma i murst-cnr project”; “study of accumulation and mixing speeds in the adriatic” prisma ii program research and experimentation for the adriatic sea; jerico european projects; i, ii, iii towards a joint european research infrastructure network for coastal observatories (20082025); vector (2004-2009): “vulnerability of italian coasts and marine ecosystems to climate change and their role in mediterranean carbon cycles”. (fisr-miur). received: 19 december 2022. accepted: 14 april 2023. publisher’s note: all claims expressed in this article are solely those of the authors and do not necessarily represent those of their affiliated organizations, or those of the publisher, the editors and the reviewers. any product that may be evaluated in this article or claim that may be made by its manufacturer is not guaranteed or endorsed by the publisher. ©copyright: the author(s), 2023 licensee pagepress, italy advances in oceanography and limnology, 2023; 14:11094 doi: 10.4081/aiol.2023.11094 this work is licensed under a creative commons attribution noncommercial 4.0 international license (cc by-nc 4.0). no nco mm er cia l u se on ly f. spagnoli and m. ravaioli32 introduction in marine environments, the fluxes of dissolved substances at the sediment-water interface (dfswis) can be due to three main factors: i) early diagenesis processes; ii) fluids coming from deeper sub-bottom layers; iii) volcanic and hydrothermal processes. early diagenesis refers to the biogeochemical processes taking place in the top centimeters of sediments (berner, 1980), where the presence of reactive organic matter (rom) provides sufficient energy for elevated microbial activity. these processes produce or consume chemicals found in pore waters (pws), thus generating concentration gradients at the sediment waterinterface that result in fluxes in or out of the sediments (figure s1). rom settling from the water column concentrates in sediment. it consists of authigenic marine particulate organic matter (mpom) and terrestrial particulate organic matter (tpom) as well as of flocs, formed by the aggregation of dissolved organic matter (dom) upon contact with the saline environment. the particulate organic matter (pom) is degraded or mineralized by microbial activity that disrupts the original pom to high molecular weight dom, which, in anaerobic environments, is in turn further hydrolyzed and fermented to monomeric low molecular weight dom (weston et al., 2006). this latter is mainly measured as dissolved organic carbon (doc). to a lesser extent, early diagenesis is also affected by processes involving the inorganic matter (im) reaching the seafloor, which consists of carbonate and silicate biogenic or lithogenic particles as well as of trace elements, pollutants, and other anthropogenic inputs (i.e. black carbon). the mineralization of pom and dom is achieved by the final electron acceptors, acting in decreasing order of free energy as follows (figure s2): dissolved oxygen (o2), nitrate (no3-), iron and manganese oxyhydroxides (fenox(oh)y, mnnox(oh)y), and sulfate (so42-). the rom mineralization induces the accumulation in pws of organic matter (om) degradation products, i.e. dissolved inorganic carbon (dic), ammonium (nh4+), phosphate (po43-), and methane (ch4), as well as the reduced forms of the electron acceptors, i.e. reduced iron and manganese (fe2+ and mn2+), dinitrogen gas (n2), and hydrogen sulfide (h2s). the intensity and type of the early diagenesis processes mainly depend on the following key factors: • reactive organic matter (rom) content. high amounts of rom, which is strongly affected by the presence of fresh om, increase the intensity of early diagenesis processes. the high reactivity of rom is due to its easy mineralization, which is particularly strong in recently produced (fresh) om, especially near the sediment-water interface (swi), where aerobic degradation is more efficient (westrich and berner, 1984). the formation of rom is favored by nutrient inputs from land, which increase primary production; high nutrient inputs are possible near the continental borders, particularly at the mouths of major rivers, which, in turn, also release large om amounts of continental origin. • fe and mn oxyhydroxides. fe and mn oxyhydroxides are important chemicals in early diagenesis processes because they can undergo dissolution or precipitation at the oxicanoxic interface in sediments. in reducing environments, fe and mn tend to dissolve from the insoluble oxidized form (fe3+, mn4+) to the soluble reduced form (fe2+, mn2+), whereas in oxidizing environments, they tend to precipitate as oxyhydroxides (figure s2, enlargement). these reactions induce an accumulation of fe and mn oxyhydroxides at the oxic-anoxic interface, albeit with a slight discrepancy between fe and mn due to their different free energy. the process is more intense where the initial amount of the oxyhydroxides of these two metals is greater and where the stability of the oxic-anoxic interface is higher, i.e. in the areas facing large river estuaries and on sea bottoms characterized by low sedimentation rates (lalond et al., 2012). • sedimentation rates. high sedimentation rates favor the incorporation of both more and less reactive om, since they result in a shorter stay of om in the oxic environment, where mineralization is faster and stronger than in anoxic environments. om preservation at greater depths induces a deepening of early diagenesis processes and a thinning of oxic and suboxic environments (figure s2). • grain size. fine grain sizes prevent oxygen penetration into sediment and promote om preservation, as do high sedimentation rates. in contrast, coarse sediments favor oxygen penetration and increase om mineralization close to the swi. • sediment reworking. the reworking of surface sediment, due to resuspension and bioturbation, plays a significant role in early diagenesis processes. an intense reworking process results in faster and stronger om decomposition; in particular, repeated om contacts with the oxidizing environment of the water column, due to resuspension, makes mineralization more efficient. moreover, dfswis can also be generated by seeps of fluids and gases, whether focused or diffused (price et al., 2015; spagnoli et al., 2020), due to hydrothermal and diagenetic processes, including low-grade metamorphism, taking place at greater sub-bottom depths (esposito et al., 2018, rovere et al., 2020). this paper addresses the dfswis in marine environments, the processes by which they are generated, and the methods by which they can be measured or calculated. it presents data regarding the dfswis measured in situ and calculated from pws in the adriatic sea, mainly by the italian consiglio nazionale delle ricerche (cnr), the university of trieste, and the slovenian marine biology station. the study focuses on research carried out from 1982 on the italian side of the central and northern adriatic sea (barbanti et al., 1995; bertuzzi et al., 1996; bertuzzi et al., 1997; cermelj et al 1997; cermelj et al., 2001; frascari et al., 1995; giordani and hammond, 1985; giordani et al., 1992; giordani et al., 1994; giordani et al., 1996; hammond et al., 1999; spagnoli and bergamini 1997; tahey et al., 1996; spagnoli et al., 2010) (figure s3, table s1). previous papers on dissolved benthic fluxes in the adriatic basin have addressed transitional ecosystems (covelli et al., 1999; covelli et al., 2011; de vittor et al., 2012; petranich et al., 2018; petranich et al, 2021), specific elements (ciceri et al., 1992; covelli et al., 2008; emili et al., 2011; emili et al., 2014; epping and helder, 1997; zago et al., 2000), or else have reworked previous data (cermelj et al., 1997, ogrinc and faganeli, 2006; testa et al., 2020) or examined other areas of the adriatic sea not considered in this paper (baric et al., 2006; giordani et al., 2002). this work addresses nutrients and dic as well as iron and manganese fluxes at the swi in the northern and centralno nco mm er cia l u se on ly dissolved fluxes of nutrients and carbon at the sediment-water interface in the adriatic sea 33 western adriatic sea. we present the results obtained by the analysis of dsfwis, measured in situ by benthic chamber deployments and on board by core incubations, and calculated from pw chemical profiles (figure s3, table s1). materials and methods benthic flux measurement dfswis can be determined in three ways (hammond et al., 1985): i) by in situ measurements using benthic chambers or eddy covariance systems; ii) by sediment core incubation in the laboratory; or iii) by modeling pw chemical composition. in particular, pw chemical concentrations can be obtained on board by pw extraction from sediment cores and subsequent microanalysis; by on-board or in situ passive sampling or microprofiling at the swi (berg et al., 2009; berner, 1980; boudreau b.p., 1996; brand et al., 2018; burdige, 2006; hammond et al., 1996; hammond et al., 1999; hammond et al., 2004; tengberg et al., 1995; viollier et al., 2003; webster et al., 1998). at present, the most reliable dfswi values are provided by the eddy covariance (an in situ approach), which integrates the dfswi over an infinite area (berg et al., 2009), whereas benthic chamber and core incubation methods measure fluxes over more limited areas and micro-profiling examines only the diffusive component of the flux at a specific point. some of these methods, as adopted and implemented by the italian cnr to measure or calculate dfswis, are described below. they include the construction and in situ deployments of benthic chambers, on-board measurements by core incubations (spagnoli f., 1994), calculation from pw chemical profiles in pw extracted on board and subsequently analyzed, and on-board micro-profiling at the swi (bergamini et al., 1997a and b; spagnoli and bergamini 1997). direct in situ measurement using benthic chambers and eddy covariance systems a benthic chamber is a device for measuring the dfswis. roughly, it is a box with a sealed top and an open bottom placed on the seafloor (figure s4a). benthic chambers range from very simple designs such as a plastic box placed on the sea or freshwater bottom environments, where water samples are collected manually, to very complex devices equipped with automatic sampling and control tools for shallow and deep aquatic environments (figure s4b-g). in italy, in situ measurements using a benthic chamber began at cnr in 1984, with a collaboration between igm-cnr (bologna, italy) and the university of southern california at los angeles (usa). this joint effort produced the first in situ measurement of dfswis in the adriatic sea (giordani and hammond et al., 1985; figure s4b). then, igm-cnr went on to build the first italian semiautomated chamber (bergamini et al., 1997 a and b; figure s4c). further research by the cnr resulted in the construction of the first italian automatic chamber (masini et al., 2001; figure s4d) and finally in the realization of the current benthic chambers and landers, respectively the adan (figure s4e) and the amerigo (spagnoli et al., 2018, 2019; figure s4f and g). benthic flux incubation experiments – by means of laboratory incubations (bertuzzi et al., 1996; bertuzzi et al., 1997) as well as in situ deployments (covelli et al., 1999) – were also carried out in the 1990s by the university of trieste in the gulf of trieste. adan is an automatic benthic chamber consisting of a polymethyl methacrylate cylinder (internal diameter, 80 cm; height, 33 cm) (figure s4e). the water sampler is managed by a commercially available tool to collect samples inside and outside the chamber or to inject tracers. the chamber is equipped with an internal rotating paddle and a multiparameter probe to measure ph, oxygen, oxidation-reduction potential, temperature, conductivity, and pressure. adan can operate on the continental shelf (down to 200 m depth) and is deployed on the seafloor through a cable to which is attached a flashing buoy for signaling and recovery. amerigo is a lander (figure s4f), that it is deployed on the bottom, operates on the seafloor, and then returns to the surface without the need for a cable, tether, or propeller. for this reason, it is equipped with ballast and buoy arrays for controlled falling, landing, and surfacing, as well as a radio, flashing light, and gps tracker for recovery. amerigo has a wide depth range, from the continental shelf to the abyssal plains (down to 6000 m depth). although its current configuration is designed to measure dfswis by means of three benthic chambers, it can also be equipped with instruments for different monitoring and measurement tasks, such as water column sensors (oxygen, ph, methane, pahs, pco2, h2s, turbidity, fluorescence) and instrumentation for seafloor measurements (micro-profilers, penetrometers, gravimeters, wave, and current meters, corers). each benthic chamber is a polymethyl methacrylate cylinder (internal diameter, 37.5 cm; height, 15 cm) with a closing lid, a rotating paddle, a device for collecting water samples inside and outside the chamber and to inject tracers in the chamber, and an oxistat device to replace oxygen in the chamber (morse et al., 1999). the hardware and software are developed in-house. once on board, the water samples collected by the adan or the amerigo are processed in an inert gas glove box and stored for subsequent analyses, which may be performed directly on board or later in the laboratory. chemical and downloaded sensor data are then used for dfswi calculation. sediment core incubation for on-board incubation, core sediments collected from the sea bottom are placed in a thermostatic bath at a constant temperature. the core tops are sealed with a lid fitted with a rotating paddle, to preserve the homogeneity of the seawater above the sediment and the thickness of the benthic boundary layer. as in a benthic chamber, water samples are collected at predetermined time intervals while physicochemical parameters are monitored by sensors (figure s4h). calculation of sediment-water interface dissolved fluxes by benthic chambers and core incubations dissolved fluxes at the swi are calculated in different ways, depending on whether they are measured by sediment incubations or calculated by modeling pw chemical composition profile. the dfswis measured by benthic chambers or core incubations are calculated by (1): no nco mm er cia l u se on ly f. spagnoli and m. ravaioli34 1. dividing the sample concentration of each chemical by the sample collection time t; 2. multiplying this gradient by the volume of the incubation and dividing it by the bottom area of the incubation chamber. namely: ji = (∂ci) (v⁄a) ∂t (1) where: ji = flux of chemical compound ith at swi; ci = concentration of chemical compound ith; t = time of sample collection or sensor measurement; v = actual volume of the incubation chamber; a = bottom area of the incubation chamber. if multiple samples have been collected inside the chamber at different times, a regression line needs to be calculated, and (1) becomes: ji = mi * h (2) where: mi = angular coefficient or slope of the regression line between ci /a and ti (figure s5); h = height of the incubation chamber. calculation of sediment-water interface dissolved fluxes from pore water chemical profiles pore water concentration profiles for dfswi calculation can be obtained in two ways: i) by on board extraction and analysis of pw from sediment cores or measured in situ by passive sampling or ii) by direct concentration measurements by a microprofiler, either in situ or on board. cnr has honed pw extraction and analysis methods, onboard micro-profiling, and in situ dialyzers. pore water extraction and analysis pore waters are extracted from a sediment core collected by a specially built corer (sw104, magagnoli & mengoli, 1995). on deck, the internal liner of the sw104 corer is placed in the extrusion frame (figure s6a). after the top of the liner has been placed in a nitrogen-filled glove box, the sediment is pushed up and cut into slices of different thicknesses (thinner near the interface and thicker at greater depths; figure s6b). after the determination of ph, redox potential, and temperature in the slices with punch-in electrodes, each slice is placed in a vessel, centrifuged (figure s6c), and then returned to the nitrogen-filled glove box. the supernatant of each slice is filtered and the resulting pw is split into aliquots, poisoned or acidified, and then stored for subsequent analyses. all operations are conducted at the same temperature as the in situ environment. for pw micro-profiling, the sediment core is placed on the micro-profiling frame; the microelectrodes (oxygen, ph, and resistivity) are lowered by a micro-manipulator (figure s7a) to obtain measurements at steps of a few microns (figure s7b), with higher resolution at the swi, which is identified by the resistivity measurements. igm-cnr has also honed a method for chemical determination of pw at the swi using in situ dialyzers (viel et al., 1991), but in the past few years this approach has been replaced by more efficient passive samplers (lei et al., 2023). diffusive flux calculation the diffusive fluxes at the swi are calculated from pw chemical profiles, by applying fick’s first law of diffusion at the concentration gradient measured at the swi (berner, 1980): ji = – ∅di (∂ci) ∂z z = 0 (3) where: ∅ = porosity di = diffusion coefficent of solute i (∂ci) ∂z z = 0 = concentration gradients at the swi of solute i. z = depth. generally, measurements of the swi concentration gradient are not very accurate, being affected by technical errors that vary depending on the method used. in particular, in the pw extraction and passive sampler methods, it is difficult to sample pw and bottom water very close to the swi, whereas in the microprofiling method the actual extension and position of the interface is difficult to determine. furthermore, dfswis include other components, besides the diffusive one, which cannot be calculated by fick’s first law due to the contribution of the biogenic irrigation and sedimentation rate (4). ftotal = fdiff + firr + fadv (4) where: ftotal = total dissolved flux at the swi; fdiff = dissolved flux due to molecular diffusion; firr = dissolved flux due to irrigation; fadv = advection of the sediment with respect to the swi. for this reason, mathematical procedures are used to calculate the interface gradient and the irrigation component, while the advective component is generally neglected (berner 1980, boudreau, 1997). from the general diagenetic equation (5): δci = ds (δ2ci) – w(1+k) (∂ci) + rz + iz = 0 (5) δt δz2 ∂z ci = concentration of the solute i z = depth r = adsorption or desorption constant or equation; i = irrigation constant or equation (non-local transport); w= burial rate k = constant as regards (5), the solution must be found on the basis of pw best fitting among the following methods (hammond et al., 1996) rz = constant c = c0 + a1z (6) rz = quadratic c = c0 + a1z + a2z2 (7) no nco mm er cia l u se on ly dissolved fluxes of nutrients and carbon at the sediment-water interface in the adriatic sea 35 rz = exponential c = c∞(1 – ebz) + c0e–bz (8) the equation that best approximates the pw profile is used to compute the gradient at the swi by calculating the first derivative at z = 0. alternatively, as a first approximation and raw calculation, the gradient can be obtained by (9) which provides difference between the concentration in the bottom water column and the concentration in the first slice (figure s8). early or raw approach: ci(bw) ci(z1) (9) where: ci(bw) = concentration of chemical compound ith in the water sample collected as close as possible to the sea bottom, or considering as bottom water the supernatant of the sediment core extruded in the nitrogen-filled glove box; ci(z1) = concentration of chemical compound i th in the first, near surface sediment core slice. comparison of measured and calculated benthic fluxes comparing measured and calculated dfswis is useful when both are available. the comparison allows to identify processes due to non-local transport or taking place at the seafloor, as well as to evaluate the accuracy of the calculation in relation to the pw analysis method applied (see section 2.5.2). comparisons of the values obtained by in situ incubation in the benthic chamber, on-board core incubation, and pw modeling at the same station in the same season usually show that the in situ method yields systematically higher fluxes than laboratory incubation, and that the latter method yields higher fluxes than pw modeling (berg et al., 2009; hammond et al., 2004). the different values produced by the three methods, particularly at the station facing the po delta mouths (station 17) are shown in figure s9. the considerable and systematically lower fluxes obtained by on-board core incubation compared with the in situ benthic chamber (figure s9, blue and light blue sticks, respectively) are attributable to the smaller area of incubated sediment core over which the fluxes are measured (berg et al., 2009; hammond et al., 2004). the smaller area limits the lateral variability of the fluxes, due particularly to the irrigation component, which increases with sediment area due to the increase of the benthic community. the fluxes determined by pw chemical profile modeling are even lower owing to two limitations: the difficulty to measure the concentration gradient as close to the swi as possible and the fact that the mathematical models applied, which consider only the diffusive component. specifically, according to the cnr approach, the gradient is calculated from the concentration in the extruded core supernatant water or in the verynear-bottom water collected by a modified niskin bottle to the concentration in the pw extracted from the first sediment core slice. these two concentrations are very close, but not precisely above and below the swi. furthermore, the concentration in the first sediment core slice is an average of the top 0.5 cm of the sediment rather than the concentration immediately below the swi, and the dz is set at 0.25 cm (the half thickness of the first sediment core slice) rather than the theoretical distance of l→0. based on the technical limitations detailed above, in situmeasured dfswis should be preferred to core incubation and pw modeling. the in situ approaches are easy to apply in shallow and very shallow marine and continental environments, because, in these conditions, deployment and recovery can be made using a cable, and the water in the chambers can be sampled using a minitube connected to a syringe, or by divers. however, the use of divers is hampered by safety restrictions, high costs as well as technical difficulties as depth increases, and becomes very problematic beyond 20-25 m. tube sampling also becomes more difficult as depth increases. on the other hand, in situ incubation at greater depths is burdened by the high costs of the instruments and the ships. this is why modeling is often preferred. results dfswi measurements taken in different seasons highlighted a variance between summer and late winter-spring (figure s10). dfswis were generally lower in winter and spring, due to both lower temperatures near the bottom and to rare primary production blooms. in summer, as temperatures rises, so do the kinetics of the biogeochemical reactions and microbial activity (hopkinson and smith, 2005; li and gregory, 1974; price and sowers, 2004; westrich and berner, 1988), therefore om mineralization and the consequent dfswis increase (westrich and berner, 1988). the same process affects primary production, which increases in spring and summer, providing further rom to the seafloor (giani et al., 2005; pugnetti et al., 2004). organic matter degradation products the dfswis of the om degradation products (nh4+, dic, po43-) declined south of the po river mouth and as far as the midadriatic depression (figure s11) along the holocene mud wedge (spagnoli et al., 2014). the dfswis were highest in front of the po river, close to the main mouth (po di pila, station 17), and just south of the po delta (stations 18, 24, 32, and e1). then they gradually declined in southward direction, at stations sharing a similar depth and distance from the coast (stations 39, 40, e3, 47) and were lowest in the mad (station 52). to the north of the po delta and in areas farther offshore with respect to the holocene muddy wedge, the fluxes were much lower than at the po mouths, except at station aa1 (in the gulf of trieste) and station 13 (facing the adige river mouth, just to the north of the po delta), where the dfswis were significantly lower than at the po river mouths (figure s12). the other open sea stations (stations 4 and 10) did not show a general trend: here, the fluxes were strongly sitedependent, low at station 4 and anomalous, even negative, at station 10 (figure s12). final electron acceptors: oxygen, nitrate, iron and manganese with regard to the final electron acceptors, dfswis showed negative values for oxygen and positive values as regard iron and manganese (figure s13). the oxygen fluxes were stronger at the stations off the po river delta, and lower southward (station 28) and in the central adriatic (station 10). nitrate dfswis were negative at all stations south of the po and positive to the north of the delta (station 13) and in the central adriatic (station 10). iron no nco mm er cia l u se on ly f. spagnoli and m. ravaioli36 and manganese dfswis declined from the po river mouths both southward and northward, although the manganese fluxes remained high to the south. discussion the strong dfswis of the om degradation products and the final electron acceptors, close to the main po river mouth (station 17), depend on the fact that this is the first deposition area of large amounts of om both of continental and marine origin (boldrin et al., 2015; frascari et al., 2006, tesi et al., 2006). this om contains highly reactive components consisting mainly of marine organisms, which proliferate in the nutrient-rich coastal waters, particularly those affected by the po river inputs. the continental om includes large amounts of low-reactive components – due to its content in om that has already undergone decomposition processes and in om from higher organisms – as well as a high rom component, consisting of recent primary production organisms from lagoons, rivers, and canals near the coast. in addition, the om settling in these areas is highly reactive because these are the first deposition areas: thus, this om has not undergone significant resuspension, which would involve some degree of mineralization and consequent loss of reactivity. another factor that contributes to the strong dfswis of the om degradation products and, to a lesser extent, of the final electron acceptors in this area, is the high sedimentation rate. this results in fast incorporation in the sediment of rom and of fe and mnoxyhydroxides, which rapidly sink to greater depths to an anoxic environment. furthermore, anoxic conditions slow down om degradation (westrich and berner 1984), enhancing its incorporation at greater depths. therefore, in this area, the dfswis of the om degradation products increased due to mineralization processes both close to the swi and to mineralization at a greater depth. this explains why the benthic fluxes were highest close to the main po river mouth (po di pila), where particle and nutrient inputs and the sedimentation rate were also highest (frignani et al., 2005). dfswis plummet a few miles offshore (figure s12, figure s13), where the western adriatic current (wac), combined with resuspension processes, removes fine sediments and rom (frascari et al., 2006). the dfswis of the om degradation products, particularly nh4+ and po43-, and of the final electron acceptors (oxygen, iron and manganese) were also high south of the po delta (figure s11, figure s13). here, the cusp of the po river delta protects the sea bottom sediments from resuspension and removal by the wac and the bora storms. as a result, rom-rich sediments settle over a wider first deposition area south of the po delta and are subject to more limited reworking. in this protected area, the high nutrient content of the po river waters also remains confined nearshore, further contributing to primary production and rom inputs to sediments. in addition, the high temperatures, calm sea, and the water column stratification characterizing this coastal area in summer, associated with high oxygen consumption, often lead to hypoxic and even anoxic conditions near the bottom (alvisi et al., 2013; djakovac et al., 2015). such conditions induce a further increase in the release of nh4+ and po43-, which do not undergo oxidation, and of fe and mn, which are subject to stronger dissolution processes also above the swi. at the southernmost stations, the dfswis of the om degradation products declined (figure s11), due to the limited rom inputs (spagnoli et al., 2014). notably, the om reaching this area has already undergone extensive transport as well as several reworking processes, which result in its being largely mineralized and less reactive. in addition, in this area, smaller amounts of nutrients limit primary production and rom deposition on the bottom. clearly, in the presence of weaker degradation processes, the final electron acceptors are less involved in the reactions, and their dfswis are also weaker. the lowest dfswi values were found in the mad. this is easily explained by the poorly reactive nature of the om reaching this area, which has largely been mineralized. this factor, combined with a lower sedimentation rate, gives rise to the setup of a plain aerobic environment in the upper few centimeters of sediment that greatly reduces the nh4+ and po43dfswis compared with the dic dfswis. this is due to the fact that nh4+ oxidizes to n2 and no3-, while po43precipitates with the feoxyhydroxides. in contrast, the po43dfswis are strongest close to the po river delta mouths, where the redox conditions promote the reduction of fe-oxyhydroxides and phosphate release. in coastal sediments north of the po river (stations 4a and 13), the fluxes of the om degradation products were weaker than in the area south of the po. this is due to low om inputs from local rivers and to the removal of part of the settled om by coastal currents and storms, which prevent a large accumulation. only in the gulf of trieste does the protected environment enable a greater accumulation of fine and om-rich sediment (ogorolec et al., 1991; ogrinc et al., 2005), hence stronger fluxes (testa et al., 2021). in the central adriatic, south of the po river delta, which receives little fine sediment and om inputs arrive (giani et al., 2003; giani et al., 2918; spagnoli et al., 2014), the dfswis were weak (figure s12). indeed, this area receives small amounts of om, which remains largely confined nearshore by the action of the wac. finally, in the center of the northern adriatic sea, which receives scanty amounts of fine sediment, om, and nutrients, the dfswis of dic and po43were low or even negative. this is due to the predominance of po43and dic precipitation and sediment incorporation, favored by aerobic environments and the carbonate sediments, over om degradation. the measurement and calculation of dfswis allowed identifying areas sharing comparable fluxes as well as biogeochemical properties (spagnoli et al, 2010; spagnoli et al., 2013, spagnoli et al., 2015), which allowed sketching a preliminary dfswi map (figure s14). these areas show different sediment characteristics due to the variables affecting its biogeochemical composition, such as: i) rom inputs (fresh marine and continental om); ii) sedimentation rate; iii) particulate and dissolved continental inputs; iv) composition of terrigenous inputs; v) distance from the main sediment sources (mainly the po river); vi) bottom sediment composition (carbonate or siliciclastic); vii) water column hydrodynamics; viii) oxygen content in near bottom waters; ix) depth. in front of the po river mouths, sediments are characterized by a diagenetic environment (type b diagenesis; figure s14) involving a high sedimentation rate and high inputs of marine rom, continental om, and fe-oxyhydroxide, which, in turn, produce high concentrations of om degradation products in pws and strong dfswis of both om degradation products and final electron acceptors. notably, these processes give rise to strongly no nco mm er cia l u se on ly dissolved fluxes of nutrients and carbon at the sediment-water interface in the adriatic sea 37 anoxic environments close to the swi. in contrast, in the mad, the greater depth and considerable distance from the main continental inputs induce a diagenetic environment characterized by highly reworked fine sediment, low sedimentation rate, low om reactivity, and mixed fine carbonate and siliciclastic sediment (type e diagenesis; figure s14). this setting is associated with major degradation at the swi, low concentrations of pw om degradation products, an oxic environment near the surface, and weak dissolved benthic fluxes. in between these two extreme types of early diagenesis, there are several environments with intermediate reactivity (types c and d, figure s14), or where coastal areas are influenced by local river inputs (type dd’, figure s14). in particular, in area c in summer, the presence of anoxic conditions in bottom waters favors greater nh4+, po43-, fe, and mn dfswis. early diagenesis plays an important role in the benthicpelagic coupling (bpc) processes, i.e. sediment processes that affect the water column and water column processes that affect sediments (testa et al., 2021). in this interaction, bathymetry plays a major role in bpc, because the same amounts of a dissolved element, released or absorbed by sediments, exert very different effects on shallow and deep-water columns. indeed, in shallow and very shallow bottoms, dissolved substances can diffuse throughout the water column up to the photic zone, thus affecting primary production; the consequent, increased nutrient supply increases the amount of rom reaching sediment, thus inducing stronger dfswis and triggers an increasingly intense cycle (joi and anderson, 2008). in the adriatic sea, the bpc is clearly strongest in the areas located in front and south of the po river delta, which receive large amounts of nutrients and rom (cozzi et al., 2019); indeed, in this area the bpc is so strong that it could be described as land-benthic-pelagic coupling. in deep marine environments, bpc processes are clearly less intense, since the substances released by the swi diffuse and dilute in a much larger volume and fail to involve the photic zone. this situation, combined with the presence in these areas of om that has lost much of its reactivity, results in a much blander bpc. the physical structure of the water column is another major factor influencing the coupling of the benthic and pelagic systems. in fact, water column stratification favors the confinement of the substances released/absorbed by sediment to the bottom layer, particularly below the pycnocline. this, in turn, boosts om mineralization, which as noted above, in the adriatic sea can induce complete anoxia of the seafloor and of bottom waters, subverting the swi biogeochemical cycles besides hampering benthic production (sturdivant et al., 2013). notably, p and metals are no longer trapped at the swi with fe and mn hydroxide precipitates, but rather are released in the water column (middelburg and levin, 1999). water column stratification also exerts effects in deeper environments, such as the mad. in this basin, dense water standing close to the bottom for one or more years undergoes an aging process characterized by gradual loss of oxygen and nutrient enrichment due to extended om mineralization. the identification of areas sharing similar biogeochemical and sedimentological composition and processes (spagnoli et al., 2014), as well as dfswis, allows providing for the first time dfswi values for dic, ammonia, and phosphates on the northern and central adriatic seafloor, enabling a more accurate definition of the budget of these elements in this ecosystem (table 1). north of the po river delta, some areas, facing local rivers, are characterized by diagenetic environments producing weak table 1. average data of dissolved fluxes at the sediment-water interface as measured and calculated in the northern and centralwestern adriatic sea (see figure s14 for location of the early diagenesis types). no nco mm er cia l u se on ly f. spagnoli and m. ravaioli38 dfswis, due to solid and dissolved inputs from the local rivers and to water column hydrodynamics, which remove and redistribute om and fine sediments along the coast and offshore (type a diagenesis; figure s14). in the gulf of trieste, bottom sediment deposition is less affected by hydrodynamic conditions (cosoli et al., 2013; pavoni et al., 2023). this results in the buildup of fine sediments and om inputs from the local rivers (ogrinc et al., 2005) and in benthic fluxes comparable to those measured south of the po river delta (type g diagenesis; figure s14). in the central adriatic the dfswis are almost null (type f diagenesis, figure s14); in some cases, where sediments are predominantly carbonate, phosphate, and dic dfswis may be negative (type f1 diagenesis; figure s14). however, since in some areas of the adriatic sea dfswi measurements have not yet been performed, the insufficient data prevent classification (figure s14: question dots). the above findings highlight the role of bottom adriatic sediments as sinks or sources of nutrients and carbon, especially as regards their contribution to the marine carbon budget. in particular, the bottom sediments in the northern and central adriatic play a large role as carbon and phosphate sinks, whereas the rest of the adriatic seafloor is a major source of nutrients, particularly of essential elements such as nitrogen and phosphorous, which control the primary production, and metals such as iron and manganese. in contrast, the coastal and fine sediments of the western adriatic are major dic sources. this dual role of bottom sediments as carbon sources and sinks should be carefully examined and considered when formulating marine carbon budgets. table 1 provides data that can be used for this purpose as well as to identify areas where data are still insufficient. another important finding regards the regeneration of nitrogen, carbon and phosphorous, which in sediments depends chiefly on om degradation and – where n and p are concerned – also on iron and oxygen. accordingly, the contribution of the bottom sediment to the recycling of these elements is the function of the main om and particulate sources. notably, among the electron acceptors, oxygen always showed negative values, thus favoring eutrophication and the setup of anoxic emergencies in the offshore areas of emilia romagna and northern marche. finally, nitrate showed both negative and positive, but altogether negligible dfswis in terms of its contribution to the water column. conclusions from 1982, several researches have been carried out in the adriatic sea by italy’s cnr and in the gulf of trieste by other research bodies. these activities allowed to develop and hone a number of approaches and methods to measure and calculate dfswis, which are described in detail in the paper. these methods include the construction and deployment of a benthic chamber, the adan, and of a lander, the amerigo, for in situ measurements; the laboratory core incubation, as well as on-board pw extraction or micro-measurements at swi, for laboratory measurements. taking measurements using these methods allowed collecting data on dfswis in the northern and central western adriatic sea and identifying the more reliable and appropriate approaches. although in situ measurements are more reliable, because of the wider area measured and of the minimal modification of the environment under measurement, laboratory core incubation, pw modeling and on-board micro-profiling are often preferred as they involve lower costs and fewer technical challenges. dfswis generally showed a seasonal trend, they are stronger in late summer and weaker in late winter, due to different near-bottom temperatures and primary production. the dfswis of the om degradation products and of the final electron acceptors commonly declined from the po river mouths southward along the holocene mud wedge. this is due to a combination of two factors: i) the increasing distance from the main particulate and nutrient sources (the po river mouths), with consequent weakening of the primary production, which is the main contributor of the highly rom to the sediment, and ii) continuous and repeated bottom sediment reworking, which reduces the rom content in surface sediments through the faster and stronger mineralization active in the aerobic water column. dfswis were weakest in the mad, where low sedimentation rate and low rom content favor the formation, near the swi, of an aerobic environment where iron, manganese and phosphorous sink into sediment. in summer, the dfswis of iron, manganese, and phosphorous were especially high in the nearshore area south of the po river delta, which is the first deposition area of the po river particulate. this process is related to the presence of calm sea conditions, water column stratification, and high oxygen consumption on the seafloor, which induce anoxic conditions hence the reduction and release of iron, manganese, and phosphates. this environmental set up highlighted the high sensitivity of the phosphorous dfswis to changes in bottom physicochemical conditions, since phosphateregeneration is strongly affected by the reduction of fe-oxyhydroxides, together with which the phosphate precipitates in oxidizing conditions. in the coastal areas to the north of the po river delta, the dfswis were weak and determined by local conditions. in particular, scarce particulate and solute inputs from the local northern rivers combined with coastal hydrodynamics, which tend to remove fine sediments and rom, result in om accumulation only in limited areas in front of the mouths of the northern rivers. in the center of the northern adriatic, dfswis were almost null and dic and phosphatedfswis were sometimes negative. this is due to the predominant carbonate and aerobic environments on the seafloor, which favors the dic and phosphorous precipitation into sediment. in the offshore areas of the central adriatic, south of the po river delta, dfswis were weak due to the poor amount of rom and fine sediments, which are confined by the wac to nearshore areas. in the nearshore areas south of the po river delta, dfswis were always considerable and were influenced by the inputs from the minor rivers. dfswis measurements and calculations in various areas of the adriatic sea allowed to identify biogeochemical facies (spagnoli et al., 2014) with comparable dfswis and to map the seafloor according to flux type and intensity. however, some areas for which data are yet to be collected remain to be classified. these findings allowed to provide, for the first time, a dfswi value that can be used for carbon and nutrient budget calculations in the adriatic sea. finally, a greater knowledge of adriatic dfswis highlighted the important role of sediments in water column no nco mm er cia l u se on ly dissolved fluxes of nutrients and carbon at the sediment-water interface in the adriatic sea 39 processes. in particular, in summer, bottom sediments play a non-negligible role in the development of hypoxic or anoxic conditions in the bottom waters of the emilia-romagna and northern marche. this is due to high oxygen consumption determined by the mineralization of the om deposited in the superficial sediments of these areas, coupled with the high water temperatures and water column stratification. furthermore, bottom sediments in the central areas of the northern adriatic sea play an important role as dic and phosphate sinks, while in the areas close to the coast, they behave like sources of carbon and phosphorous. this dual role of the adriatic sea sediments should be carefully considered in carbon and nutrient budget modeling, especially when assessing the contribution of dic to carbon fluxes at the atmosphere-sea interface and to seawater acidification. in this context, additional data about dfswis are needed to assess in greater detail the contribution of sediments to the marine budget of the main chemical compounds that play a role in global climate and marine cycles changes. finally, as noted by testa et al. (2021), most recent efforts directed at the quantification of the nutrient and carbon budgets in the adriatic sea have been devoted to developing numerical modeling, not collecting fresh data (alvarez et al., 2022). yet, new information would be extremely useful, especially at a time when far-reaching global changes – particularly increased temperature and solid and liquid riverine variations of contributions – are affecting marine biogeochemical cycles in the adriatic sea. references alvisi f, giani m, ravaioli m, giordano p, 2013. role of sedimentary environment in the development of hypoxia and anoxia in the nw adriatic shelf (italy). estuarine, coastal and shelf science. 128:9-21. artegiani a, paschini e, russo a, et al., 1997. the adriatic sea general circulation. part i: air–sea interactions and water mass structure. journal of physical oceanography. 27:1492-514. barbanti a, bergamini mc, frascari f, et al., 1995. diagenetic processes and nutrient fluxes at the sediment water interface, northern adriatic sea, italy, marine & freshwater research. 46:55-67. baric a, kuspilic g, matijevic s, 2002. nutrient (n, p, si) fluxes between marine sediments and water column in coastal and open adriatic hydrobiologia. 475/476:151–9. bergamini mc, frascari f, marcaccio m, et al., 1997a. flussi bentici. relazione finale per il progetto prisma i, sotto progetto “flussi da e verso i fondali”. istituto di geologia marina cnr, bologna, italy: 7 pp. bergamini mc, frascari f, marcaccio m, et al., 1997b. processi diagenetici. relazione finale per il progetto prisma i, sotto progetto “flussi da e verso i fondali”. istituto di geologia marina cnr, bologna, italy: 9 pp. berelson wm, hammond de, 1986. the calibration of a new freevehicle benthic flux chamber for use in the deep sea. deep sea research part a. oceanog. research papers. 33:1439-54. berg p, glud rn, hume a, et al., 2009. eddy correlation measurements of oxygen uptake in deep ocean sediments. limnol. and oceanog. methods. 7:576-84. berner ra, 1980. early diagenesis: a theoretical approach. no. 1. princeton university press, princeton, usa. 256 pp. bertuzzi a, faganeli j, brambati a, 1996. annual variation of benthic nutrient fluxes in shallow coastal waters (gulf of trieste, northern adriatic sea). mar. ecol. 17:261-78. bertuzzi a., faganeli j, welker c, brambati a, 1997. benthic fluxes of dissolved inorganic carbon, nutrients and oxygen in the gulf of trieste (northern adriatic) water, air, and soil pollution. 99:305–14. boldrin a, langone l, miserocchi s, et al., 2005. po river plume on the adriatic continental shelf: dispersion and sedimentation of dissolved and suspended matter during different river discharge rates. marine geology. 222– 223:135–58. bortoluzzi g, spagnoli f, aliani s, et al., 2014. new geological, geophysical and biological insights on the hydrothermal system of the panarea – basiluzzo volcanic complex (aeolian islands, tyrrhenian sea. sgi-simp 2014. milan, italy. boudreau bp, 1997. diagenetic models and their implementation: modelling transport and reactions in aquatic sediments. springer, new york, usa. 428 pp. burdige d j, 2006. geochemistry of marine sediments. princeton university press, princeton, usa. brand a, müller b, wüest a, et al., 2007. microsensor for in situ flow measurements in benthic boundary layers at submillimeter resolution with extremely slow flow. limnology and oceanography: methods. 5:185-91. cermelj b, bertuzzi a, faganeli j, 1997. modelling of pore water nutrient distribution and benthic fluxes in shallow coastal waters gulf of trieste, northern adriatic. water air soil pollut. 99:435-44. cermelj b, ogrinc n, faganeli j, 2001. anoxic mineralization of biogenic debris in near-shore marine sediments gulf of trieste, northern adriatic, the science of the total environment. 266:143-52. ciceri g, maran c, martinotti w, queirazza,g, 1992. geochemical cycling of heavy metals in a marine coastal area: benthic flux determination from pore water profiles and in situ measurements using benthic chambers. hydrobiologia. 235: 501-17. cosoli s, ličer m, vodopivec m, malačič v, 2013. surface circulation in the gulf of trieste (northern adriatic sea) from radar, model, and adcp comparisons. journal of geophysical research: oceans. 118:6183-200. covelli s, emili a, acquavita a, et al., 2011. benthic biogeochemical cycling of mercury in two contaminated northern adriatic coastal lagoons continental shelf research. 31:1777-89. covelli s, faganeli j, de vittor c, et al., 2008. benthic fluxes of mercury species in a lagoon environment (grado lagoon, northern adriatic sea, italy) applied geochemistry. 23: 529-46 covelli s, faganeli j, horvat m, brambati a, 1999. porewater distribution and benthic flux measurements of mercury and methylmercury in the gulf of trieste (northern adriatic sea) estuarine, coastal and shelf science. 48:415-28. cozzi s, ibáñez c, lazar l, et al., 2019. flow regime and nutrient-loading trends from the largest south european watersheds: implications for the productivity of mediterranean and black sea’s coastal areas, water. 11:1-27. no nco mm er cia l u se on ly f. spagnoli and m. ravaioli40 de vittor c, faganeli j, emili a, et al., 2012. benthic fluxes of oxygen, carbon and nutrients in the marano and grado lagoon (northern adriatic sea, italy) estuarine, coastal and shelf science. 113:57-70. djakovac t, supić n, aubry fb, et al., 2015. mechanisms of hypoxia frequency changes in the northern adriatic sea during the period 1972–2012. journal of marine systems. 141:179-89. emili a, carrasco l, acquavita a, covelli s, 2014. a laboratoryincubated redox oscillation experiment to investigate hg fluxes from highly contaminated coastal marine sediments (gulf of trieste, northern adriatic sea) environmental science and pollution research. 21:4124-33. emili a, koron n, covelli s, et al., 2011. does anoxia affect mercury cycling at the sediment–water interface in the gulf of trieste (northern adriatic sea)? incubation experiments using benthic flux chambers applied geochemistry. 26:194-204. epping eh, helder w, 1997. oxygen budgets calculated fromin situ oxygen microprofiles for northern adriatic sediments. continental shelf research. 17:1737-64. esposito v, andaloro f, canese s, et al., 2018. exceptional discovery of a shallow-water hydrothermal site in the sw area of basiluzzo islet (aeolian archipelago, south tyrrhenian sea): an environment to preserve. plos one. 13:e0190710. frignani m, langone l, ravaioli m, et al., 2005. fine�sediment mass balance in the western adriatic continental shelf over a century time scale. marine geology. 222:113–33. frascari f, bergamini mc, spagnoli f, marcaccio m, 1995. effetti sull’ambiente marino della piattaforma antares. indagini sui processi geochimico-sedimentologici. geda/ambi agip, milano (italy), ceom-doc a160-011. frascari f, spagnoli f, marcaccio m, giordano p, 2006. anomalous po river flood event effects on sediments and the water column of the northwestern adriatic sea. climate research. 31:151-65. froelich p, klinkhammer gp, bender ml, luedtke n a, heath g r, cullen d, maynard v, 1979. early oxidation of organic matter in pelagic sediments of the eastern equatorial atlantic: suboxic diagenesis. geochimica et cosmochimica acta, 43(7), 1075-1090. giani m, berto d, rampazzo f, et al., (2009). origin of sedimentary organic matter in the north�western adriatic sea. estuarine, coastal and shelf science. 84:573–83. giani m, savelli f, boldrin a, 2003. temporal variability of particulate organic carbon, nitrogen and phosphorus in the northern adriatic sea. hydrobiologia. 494:319–25. giordani p, balboni v, hammond de, et al., 1996. benthic mineralization: temporal variability and coupling to water column protiction. in: price n.b. (editor) euromargeas final report, mast ii programme, eu, pp. 208-218. giordani p, balboni v, miserocchi s, et al., 1992. step, adria cruise. cruise report. igm-cnr, pp. 46. giordani p, donnelly a, herbert ra, seritti a, balboni v, del vecchio r, miserocchi s, sorgente d, vitale f, nannicini l, 1996a. assessing diagenetic efficiency in biogeochemical cycling: results from laboratory incubation experiments. in: price n.b. (editor) euromarge-as final report, mast ii programme, eu, 208-218. giordani p, hammond de, 1985. techniques for measuring benthic fluxes of 222rn and nutrients in coastal waters. technical report n.20, igm-cnr, pp. 33. giordani p, hammond de, berelson wm, et al., 1992. benthic fluxes and nutrient budgets for sediments in the northern adriatic sea: burial and recycling efficiencies. in marine coastal eutrophication (pp. 251-275). elsevier, amsterdam, the netherlands. giordani p, miserocchi s, balboni v, et al., 1994. emas 94 cruise report. results presented at the mast workshop, barcellona. technical report n. 30. igm-cnr bologna, italy. hammond de, cummins km, mcmanus j, et al., 2004. methods for measuring benthic nutrient flux on the california margin: comparing shipboard core incubations to in situ lander results. limnol. and oceanog.: methods. 2:146-59. hammond de, fuller c, harmon d, et al., 1985. benthic fluxes in san francisco bay. hydrobiologia. 129:69-90. hammond de, giordani p, berelson wm, poletti r, 1999. diagenesis of carbon and nutrients and benthic exchange in sediments of the northern adriatic sea. marine chemistry. 66:53-79. hammond de, mcmanus j, berelson wm, 1997. early diagenesis of organic material in equatorial pacific sediments. stoichiometry and kinetics oceanographic literature review. 5:458. hassoun aer, bantelman a, canu d, et al., 2022. ocean acidification research in the mediterranean sea: status, trends and next steps. front. mar. sci. 9. hopkinson jr. cs, smith em, 2005. estuarine respiration: an overview of benthic, pelagic, and whole system respiration. in p.a. del giorgio, p.j. williams (eds.), respiration in aquatic systems (pp. 122–146). oxford, uk: oxford university press. joye sb, anderson ic, 2008. nitrogen cycling in coastal sediments. in d.g. capone, d.a. bronk, m.r. mulholland, & e.j. carpenter (eds.), nitrogen in the marine environment (2nd edn., pp. 867–915). amsterdam, the netherlands: elsevier. lalonde k, alfonso m, alexandre o, ge´linas y, 2012. preservation of organic matter in sediments promoted by iron. nature. 483:198-200. lei p, chen m, rong n, et al., 2023. a passive sampler for synchronously measuring inorganic and organic pollutants in sediment porewater: configuration and field application. journal of environm. sciences. 136:201-12. li y, gregory s, 1974. diffusion of ions in sea-water and in deep-sea sediments. geoch. cosmoch. acta. 38:703-14. magagnoli a, mengoli m, 1995. carotiere a gravità sw-104. rapporto tecnico cnr 27.pp. 45. milucka j, ferdelman tg, polerecky l, et al., 2012. zero-valent sulphur is a key intermediate in marine methane oxidation. nature. 491:541-6. middelburg jj, levin la, 2009. coastal hypoxia and sediment biogeochemistry. biogeosciences. 6:1273–93. morse jw, boland g, rowe gt, 1999. a gilled benthic chamber for extended measurement of sediment-water fluxes. marine chemistry. 66:225-30. ogorelec b, misic m, faganeli j, 1991. marine geology of the no nco mm er cia l u se on ly dissolved fluxes of nutrients and carbon at the sediment-water interface in the adriatic sea 41 gulf of trieste (northern adriatic): sedimentological aspects. marine geology. 99:79–92. ogrinc n, faganeli j, 2006. phosphorus regeneration and burial in near-shore marine sediments (the gulf of trieste, northern adriatic sea) estuarine, coastal and shelf science. 67: 579-88. ogrinc n, fontolan g, faganeli j, covell s, 2005. carbon and nitrogen isotope compositions of organic matter in coastal marine sediments (the gulf of trieste, n adriatic sea): indicators of sources and preservation. marine chemistry. 95:163-81. pavoni e, petranich e, signore s, et al., 2023. fluxes of settling sediment particles and associated mercury in a coastal environment contaminated by past mining (gulf of trieste, northern adriatic sea). journal of soils and sediments. 2023:1-12. petranich e, covelli s, acquavita a, et al., 2018. benthic nutrient cycling at the sediment-water interface in a lagoon fish farming system (northern adriatic sea, italy) science of the total environment. 644:137-49. petranich e, crosera m, pavoni e, et al., 2021. behaviour of metal(loid)s at the sediment-water interface in an aquaculture lagoon environment (grado lagoon, northern adriatic sea, italy). appl. sci. 11:2350. price pb, sowers t, 2004. temperature dependence of metabolic rates for microbial growth, maintenance, and survival. proceedings of the national academy of sciences. 101:4631-6. price re, larowe de, italiano f, et al., 2015. subsurface hydrothermal processes and the bioenergetics of chemolithoautotrophy at the shallow-sea vents off panarea island (italy). chemical geology. 407:21-45. pugnetti a, acri f, alberighi l, et al., 2004. phytoplankton photosynthetic activity and growth rates in the nw adriatic sea. chemistry and ecology. 20:399–409. sturdivant sk, brush mj, diaz rj, 2013. modeling the effect of hypoxia on macrobenthos production in the lower rappahannock river, chesapeake bay, usa. plos one. 8:e84140. rovere m, mercorella a, frapiccini e, et al., 2020. geochemical and geophysical monitoring of hydrocarbon seepage in the adriatic sea. sensors. 20:1504. savini a, malinverno e, etiope g, et al., 2009 shallow seeprelated seafloor features along the malta plateau (sicily channel-mediterranean sea): morphologies and geoenvironmental control of their distribution. marine and petroleum geology. 26:1831-48. spagnoli f, 1994. diagenesi precoce e processi di scambio tra acqua e sedimento in condizione di quiete e di risospensione di un’area marina a sud del delta del po. università degli studi di trieste, tesi di dottorato di ricerca in scienze ambientali (oceanografia geologica e risorse), v ciclo, 125 pp. spagnoli f, bergamini mc, 1997. water-sediment exchange of nutrients during early diagenesis and resuspension of anoxic sediments from the northern adriatic sea shelf. water, air and soil poll. 99:541-56. spagnoli f, dell’anno a, de marco a, et al., 2010 biogeochemistry, grain size and mineralogy of the central and southern adriatic sea sediments: a review. chemistry and ecology. 26:19-44. spagnoli f, de marco r, giuliani g, et al., 2020. dissolved heavy metal fluxes at sediment-water interface in polluted sediments of the adriatic sea, egu general assembly 2020. egu2020-19366. spagnoli f, dinelli e, giordano p, et al., 2014. sedimentological, biogeochemical and mineralogical facies of northern and central western adriatic sea. journal of marine systems. 139:183-203. spagnoli f, gabriella b, capaccioni b, giordano p, 2013. benthic nutrient fluxes in central and southern adriatic and ionian seas. 40th ciesm congress proceedings. rapp. comm. int. mer médit. 40:266. spagnoli f, giordano p, ravaioli m, 2018. marine dissolved carbon fluxes at sediment-water interface and early diagenesis processes. libro degli abstract. 1° workshop nazionale: terra, vita e clima il ciclo del carbonio. area della ricerca cnr pisa, 22 23 novembre 2018, p. 65. spagnoli f, giuliani g, martinotti v, et al., 2018. amerigo and cba: a new lander and a new automatic benthic chamber for dissolved benthic flux measurements. in 2018 ieee international workshop on metrology for the sea; learning to measure sea health parameters (metrosea) (pp. 206-211). spagnoli f, kaberi h, giordano p, et al., 2015. benthic fluxes of dissolved heavy metals in polluted sediments of the adriatic sea. scientific conference proceedings. scientific conference: integrated marine research in the mediterranean and the black sea. e. papathanassiou, n. streftaris & l. giannoudi, editors, pp. 301-302. 399 pp. spagnoli f, penna p, giuliani g, et al., 2019. the amerigo lander and the automatic benthic chamber (cba): two new instruments to measure benthic fluxes of dissolved chemical species. sensors, 19(11), 2632. pp. 1-30. tahey tm, duineveld gca, dewilde pawj, et al., 1996. sediment o-2 demand, density and biomass of the benthos and phytopigments along the northwestern adriatic coast: the extent of po enrichment. oceanologica acta. 19: 117-30. tengberg a, de bovee f, hall p, et al., 1995. benthic chamber and profiling landers in oceanography a reviewof design, technical solutions and functioning. prog. oceanog. 35: 253-94. tesi t, miserocchi s, langone l, et al., 2006. sources, fate and distribution of organic matter on the western adriatic continental shelf, italy. water, air, and soil pollution: focus. 6:593–603. testa j, faganeli j, giani m, et al., 2021. advances in our understanding of pelagic-benthic coupling. in: malone t., malej a., faganeli j., (eds) coastal ecosystems in transition: a comparative analysis of the northern adriatic and chesapeake bay, geophysical monograph. 256:147-75. viel m, barbanti a, langone l, et al., 1991. nutrient profiles in the pore water of a deltaic lagoon: methodological considerations and evaluation of benthic fluxes. estuarine, coastal and shelf science, 33:361-82. viollier e, rabouille c, apitz se, et al., 2003. benthic biogeochemistry: state of the art technologies and guidelines for the future of in situ survey. j. exp. mar. biol. ecol. 285:5-31. no nco mm er cia l u se on ly f. spagnoli and m. ravaioli42 webster it, teasdalev pr, grigg nj, 1998. theoretical and experimental analysis of peeper equilibration dynamics. environ. sci. technol. 32:1727–33. weston nb, porubsky wp, samarkin va, et al., 2006. porewater stoichiometry of terminal metabolic products, sulfate, and dissolved organic carbon and nitrogen in estuarine intertidal creek-bank sediments. biogeochemistry. 77: 375-408. westrich jt, berner ra, 1984. the role of sedimentary organic matter in bacterial sulfate reduction: the g model tested 1. limnology and oceanog. 29:236-49. westrich jt, berner r a, 1988. the effect of temperature on rates of sulfate reduction in marine sediments. geomicrobiology journal. 6:99-117. zago c, capodaglio g, ceradini s, et al., 2000. benthic fluxes of cadmium, lead, copper and nitrogen species in the northern adriatic sea in front of the river po outflow, italy. science of the total environ. 246:121-37. online supplementary material: figure s1. schematic view of early diagenesis processes. figure s2. early diagenesis processes. a) schematic drawing of dissolved ions produced and/or consumed in pore waters by early diagenesis processes (modified from froelich et al., 1979; milucka et al., 2012). enlargement: fe and mn oxyhydroxide cycles; b) biogeochemical reactions taking place in the top few centimeters of sediment (blue: final electron acceptors; red: organic matter degradation products). figure s3. stations in the central and northern adriatic sea where in situ measurements, on-board core incubations, and pore water chemical profile modeling were conducted from 1982. symbols, season, and depth color codes are reported in the legend, stations are listed in table s1. figure s4. a) schematic drawing of a benthic chamber (spagnoli et al., 2018, 2019). bc, benthic chamber: 1) device for collecting water samples inside and outside the chamber or to inject tracers into the chamber (vampire); 2) multiparameter probe (hydrolab ms5) to measure oxygen, ph, eh, temperature, conductivity (i.e. salinity) and pressure (depth) in the chamber; 3) housing of the electronics (idec microsmart fc6a plc); 4) housings of batteries; 5) rotating paddle; 6) motors. b-g) evolution of the benthic chamber technology at the italian cnr: b) benthic lander deployed on the adriatic seafloor in 1982 (berelson and hammond, 1986) (credit, gabriele marozzi); c) first manual benthic chambers involving sampling by divers; d) first automatic benthic chamber developed at igm-cnr of bologna (masini et al., 2001); e) the adan chamber (spagnoli et al., 2018, 2019). f) the amerigo lander on deck (spagnoli et al., 2018, 2019); g) the amerigo deployed on the seafloor. h) incubation of cores for the measurement of dissolved fluxes at the sediment-water interface in a thermostatic box (spagnoli f., 1994). figure s5. concentrations of chemical compound ith versus time; m is the angular coefficient of the straight line calculated by the minimum square method. figure s6. pictures of some steps of the collection of pore waters from sediment cores by the sw104 corer. a) the sediment core mounted on the extrusion frame with its top in the inert gas glove box; b) subsampling of the sediment core slices; c) centrifugation of sediment slices d) pore water filtration, aliquot preparation poisoning and acidification in the inert gas glove box. figure s7. on-board profiling at the sediment-water interface on a core by a microelectrode (cnr, cruise vector 632). a) micro-profiling apparatus; b) drawing of the results (o2 concentrations). figure s8. pore water and bottom water concentration profile of chemical compound ith obtained from pore water extraction. figure s9. dfswis of nh4+ and po43measured by in situ benthic chamber deployments (blue), by on-board core incubation (clear blue), and by pore water modeling (grey) at the same station in the same season. see figure s3 for station locations. figure s10. variability of nh4+ and po43dfswis in different seasons (summer and spring) at stations located in the same area (in this case, facing the po di pila mouth). see figure s3 for station locations. figure s11. dfswis of organic matter degradation products (nh4+, dic, po43-) at stations located along the holocene mud wedge on the adriatic seafloor, between the po river delta and the mad, in late summer. blue: in situ deployments, light blue: on-board incubations, grey: pore water calculations. pw, pore water. figure s12. dfswis of organic matter degradation products (nh4+, dic, po43-) on the adriatic seafloor north of the po delta and offshore in the central adriatic. pw, pore water. figure s13. dfswis of the final electron acceptors (o2, no3-, fe, mn) at the station established on the adriatic seafloor. pw, pore water. figure s14. areas characterized by different early diagenesis processes and flux intensities at the water-sediment interface in the adriatic sea. question marks indicate areas where the available data are insufficient for classification. table s1. dissolved fluxes at the sediment-water interface as measured and calculated from pore water profiles in the northern and central-western adriatic sea. units: mmol*m-2*d-1. no nco mm er cia l u se on ly layout 1 advances in oceanography and limnology, 2015; 6(1/2): 1 editorial doi: 10.4081/aiol.2015.5684 with this first double issue of 2015, advances in oceanography and limnology (aiol journal) is facing an important regime shift. published in house until 2007 by the italian association of oceanography and limnology (aiol, www.aiol.info) as the proceedings of the aiol national congress, it has been transformed starting from 2011 in a regular scientific journal with the aim of offering a new publication outlet for oceanographers and limnologists. after five volumes (ten issues) published under the editorial responsibility of the former aiol president roberto danovaro between 2011 and 2014, the journal, still sponsored by aiol, has now passed to pagepress (www.pagepress.org). in september 2015, the presidency council of the aiol, corroborated by the vote of the aiol members assembly held in verbania (italy), has passed to us the editorial responsibility of the journal. we have enthusiastically accepted this challenge and we would like here to briefly explain how we intend to pave the way for a new deal of the aiol journal. the aiol journal was funded as an interdisciplinary journal, and we want to uphold its wide scope, which embraces both fundamental and applied oceanography and limnology, with focus on both single and multiple disciplines. in the modern scientific vision, the adoption of a multidisciplinary approach is essential for understanding the mechanisms of ecosystem functioning and their alteration caused by natural and anthropogenic influences. such a task requires a diversified editorial board, with associate editors specialized in different aquatic scientific disciplines, and yet strongly motivated to put together the different pieces that represent the core of the interdisciplinary approach. at present, the aiol journal relies on twin editors in chief, while a new, compact and agile group of associate editors will be added in the near future. the correct identification of referees is essential for an accurate evaluation of the submitted contributions. we are strongly aware that the selfless and invaluable scientific work of qualified referees are essential for the final quality and impact of the published papers and for the reputation of the aiol journal. with the 2015 issue, the aiol journal is now published as a new open access, peer-reviewed journal. in the intricate world of publishers, we admit we cannot now compete with highly reputed journal with a longer tradition of publication. we also know that this world is increasingly being populated by a plethora of open access journals from any part of the globe, which increases a lot the competition for attracting esteemed authors as well as receiving high quality papers. so, why the need of another open access journal? the response is astonishingly simple: our aim is to provide a new open-access, free of charge floor for the large community of world oceanographers and freshwater scientists, with an eye on the mediterranean community. this ambitious aim is possible by the free dedication and support of the italian association of oceanography and limnology. wide space is given to regular articles, review, short notes and opinion papers. proceedings that focus on topics that are timely and of interest to a significant number of aquatic scientists will be equally evaluated for their inclusion in specific issues. the publishing house is guaranteeing a fast and high quality editorial processing of the submitted papers. with this last issue, about 4 months were required from the date of submission and the online publication of papers. given its present infancy stage, the aiol journal is not yet fully indexed. the aiol presidency council in charge, of which we are honoured to be part of, fostered the publication of the 2015 issue under pagepress in order to save the publishing effort paid during the first four years of the journal’s life and ensure indexing agencies to follow up the journal. this will give advances in oceanography and limnology an opportunity to be monitored and, in a hopefully near future, to be ultimately ranked. we are conscious that making the aiol journal attractive for highly reputed authors will be the major part of the challenge. we will put a special effort to select high quality papers through a rigorous peer reviewing process, also profiting of the future board of international associate editors. we will also put a special effort for mining thematic reviews from esteemed oceanographers and limnologists all over the world, providing a comprehensive state of the art overview and, at the same time, indicating future directions and perspectives in a given field. the ultimate undisguised ambition is thus to give the aiol journal an international position in the wide panorama of the already available scientific literature with “water-oriented” aims and scope. antonio pusceddu nico salmaso editors, advances in oceanography and limnology a new deal for advances in oceanography and limnology (aiol journal) antonio pusceddu,1 nico salmaso2 1dipartimento di scienze della vita e dell’ambiente, università degli studi di cagliari, 09126 cagliari, italy; 2dipartimento agroecosistemi sostenibili e biorisorse, fondazione e. mach, san michele all’adige, trento, italy corresponding author: apusceddu@unica.it ; nico.salmaso@fmach.it advances in oceanography and limnology layout 1 handbook of cyanobacterial monitoring and cyanotoxin analysis j. meriluoto, l. spoof, g.a. codd (eds.), 2017 john wiley & sons ltd., west sussex, uk isbn: 978-1-119-06868-6 pages: 576; € 168,00 cyanobacteria are impacting widely on the daily activities of the general public. they are increasingly recognized as agents having possible adverse effects on water resources, on health of human and/or animals and therefore should be monitored and managed. to fill existing gaps this valuable handbook which is edited by three highly respected experts in the areas microbiology, algae biochemistry, metabolomics, and environmental science provide reviews, practical methods and standard operating procedures. scientists from both university and industry contributed a series of chapters summarizing and introducing the reader to cyanobacteria (section i and ii), reviewing sampling, monitoring including risk management (section iii), and toxins plus other bioactive compounds in cyanobacteria (section iv). these sections are followed by a comprehensive survey of state-of-the-art analytical methods to characterize different metabolites in cyanobacteria (section v) and methodological considerations (section vi). section vii gives several standard operating procedures around various aspects of cyanobacteria analysis. the style of this book is not only informative, but also a joy to read. it is well produced and illustrations are given in a very adequate way including a bibliography following each chapter. therefore, the book should be part of any library of cyanobacteria microbiologist and cyanotoxin chemist. furthermore, the book provides very interesting material to be purchased by those with a much broader interest in algae chemistry and ecology. stefan martens food quality and nutrition department, research and innovation centre fondazione edmund mach, san michele all’adige, italy advances in oceanography and limnology, 2017; 8(2): 242 book review doi: 10.4081/aiol.2017.7221 this work is licensed under a creative commons attribution-noncommercial 4.0 international license (cc by-nc 4.0). no nco mm er cia l u se on ly layout 1 article erratum macrobenthos of lagoon ecosystems: a comparison in vegetated and bare sediments paolo magni1,*, maria flavia gravina1,2,3 1consiglio nazionale delle ricerche, istituto per lo studio degli impatti antropici e sostenibilità in ambiente marino (cnr-ias), località sa mardini, torregrande, oristano, italy; 2dipartimento di biologia, università di roma “tor vergata”, rome, italy; 3consorzio nazionale interuniversitario per le scienze del mare (conisma), rome, italy abstract this corrects the article published on advances in oceanography and limnology 2023;14:11124 doi: https://doi.org/10.4081/aiol.2023.11124 table 1. list of the top discriminating taxa with contribution more than 5% by simper analysis at the three investigated areas, c. nodosa-dominated (cym), mixed-macrophyte/detritus (mix) and unvegetated (unv). species category: o, opportunistic; c, common in coastal sheltered waters; b, brackish; m, properly marine. cym vs mix average contribution cumulative average average dissimilarity % % abundance abundance cym mix cerithium lividulum m 8.39 13.71 13.71 7.77 3.88 microdeutopus gryllotalpa c 8.36 13.67 27.37 6.46 2.23 loripes orbiculatus c 6.98 11.41 38.78 3.83 2.40 gammarus aequicauda c 6.50 10.62 49.41 3.50 2.33 abra tenuis m 4.74 7.74 57.15 1.63 1.67 sphaeroma serratum m 4.60 7.52 64.67 4.00 0.25 nainereis laevigata c 3.93 6.42 71.08 2.19 0.56 perinereis rullieri m 3.66 5.98 77.06 2.06 3.29 abra segmentum c 3.59 5.86 82.92 1.08 2.15 chironomidae o 2.11 3.45 86.37 0.19 1.04 tanais dulongii o 1.33 2.17 88.54 1.15 0.06 cyathura carinata b 0.94 1.53 90.07 0.73 0.00 gibbula adansoni m 0.85 1.39 91.46 0.08 0.42 cerastoderma glaucum b 0.79 1.30 92.76 0.23 0.35 phylo norvegicus c 0.61 1.00 93.76 0.65 0.02 dispio uncinata m 0.56 0.91 94.67 0.00 0.25 pirenella conica b 0.51 0.84 95.51 0.00 0.27 cym vs unv average contribution cumulative average average dissimilarity % % abundance abundance cym unv abra tenuis m 26.91 27.93 27.93 1.63 32.40 chironomidae o 16.39 17.01 44.94 0.19 16.00 capitella capitata o 8.23 8.54 53.48 0.02 10.30 hydrobia acuta b 8.15 8.46 61.94 0.00 10.30 cerithium lividulum m 8.11 8.42 70.36 7.77 0.00 cerastoderma glaucum b 6.08 6.31 76.67 0.23 8.33 microdeutopus gryllotalpa c 5.50 5.71 82.38 6.46 0.13 loripes orbiculatus c 3.17 3.29 85.68 3.83 0.00 sphaeroma serratum m 2.95 3.06 88.74 4.00 0.00 gammarus aequicauda c 2.64 2.74 91.47 3.50 0.17 perinereis rullieri m 2.34 2.43 93.90 2.06 0.00 nainereis laevigata c 2.05 2.13 96.03 2.19 0.00 mix vs unv average contribution cumulative average average dissimilarity % % abundance abundance mix unv abra tenuis m 28.62 30.90 30.90 1.67 32.40 chironomidae o 17.80 19.22 50.13 1.04 16.00 hydrobia acuta b 8.86 9.57 59.70 0.00 10.30 capitella capitata o 8.72 9.41 69.11 0.19 10.30 cerastoderma glaucum b 6.51 7.03 76.14 0.35 8.33 cerithium lividulum m 4.75 5.13 81.27 3.88 0.00 perinereis rullieri m 4.00 4.32 85.59 3.29 0.00 gammarus aequicauda c 2.81 3.03 88.62 2.33 0.17 abra segmentum c 2.72 2.94 91.56 2.15 0.00 microdeutopus gryllotalpa c 2.57 2.78 94.34 2.23 0.13 loripes orbiculatus c 2.32 2.51 96.85 2.40 0.00 no nco mm er cia l u se on ly layout 1 cost (european cooperation in science and technology) is a funding agency for research and innovation networks. cost actions help connect research initiatives across europe and enable scientists to grow their ideas by sharing them with their peers. this boosts their research, career and innovation. www.cost.eu this themed issue results from international collaboration within cost action es1105 “cyanocost cyanobacterial blooms and toxins in water resources: occurrence, impacts and management” (www.cyanocost.net) running from 2012 to 2016. the cyanocost action is acknowledged for adding value to this work through networking and knowledge sharing with european experts. the action has involved 32 countries. state of the art research and management capabilities in europe on cyanobacteria have benefited from input from the basic and applied life sciences, the human and animal health sectors, water engineers, economists and planners. many of these professional groups have been brought together and they interacted favourably within the framework of cyanocost. the general goals of the action have been to widen awareness, spread relevant technical competence, and share risk management experience and expertise in the field. the action has developed and provided tools to end-users (public health and environment authorities, water utilities, aquaculture, tourism and recreation sectors) by pooling and coordinating expertise from throughout europe and has contributed to harmonizing methods and practices, thereby protecting public health, enterprises and investments. the 13 papers published in this themed issue are grouped under six thematic topics and the contents of the papers are briefly explained below. the descriptions of the papers have been kindly compiled by the guest editors. cyanobacteria occurrence rare occurrence of nine microcystis species (chroococcales, cyanobacteria) in a single lake (lake dojran, fyr macedonia) by krstić, aleksovski and komárek presents an ecological and thorough taxonomic study of the plankton community in lake dojran. it revealed the co-existence of nine microcystis species, provided detailed morphological features, and corroborated the necessity to change the accepted morphospecies concept into a separation of microcystis taxa as distinct species. cyanobacteria and cyanotoxin environmental occurrence and monitoring a comparative study of the metabolic profiles of common nuisance cyanobacteria in southern perialpine lakes by cerasino, capelli and salmaso used target and non-target metabolite profiling to identify differences in the production of known cyanotoxins and other secondary metabolites, primarily non-ribosomal peptides, in 14 strains of 5 cyanobacterial species from several perialpine lakes. monitoring a newly re-born patient: water quality and cyanotoxin occurrence in a reconstructed shallow mediterranean lake by gkelis, panou, chronis, zervou, christophoridis, manolidi, triantis, kaloudis, hiskia, kagalou and lazaridou discusses cyanobacterial abundance and diversity, toxin levels, physico-chemical and ecological characteristics in the recently reconstructed lake karla in greece, showing immediately occurring problems with cyanobacterial water blooms, toxin production and degradation of ecological status. molecular detection of hepatotoxic cyanobacteria in inland water bodies of the marmara region, turkey by köker, akçclaan-albay, albay and neilan for the first time identifies production of the cyanobacterial hepatotoxins microcystin (mc) and nodularin (nod) using pcr in combination with hplc in bloom samples and cyanobacterial strains isolated from lakes in marmara. advances in oceanography and limnology, 2017; 8(1): 1-3 article doi: 10.4081/aiol.2017.6674 this work is licensed under a creative commons attribution-noncommercial 4.0 international license (cc by-nc 4.0). foreword to the themed issue “cyanobacteria” triantafyllos kaloudis,1 jussi meriluoto,2,3* ludek blaha4 1department of water quality control, athens water supply and sewerage company, athens, greece; 2biochemistry, faculty of science and engineering, åbo akademi university, tykistökatu 6a, 20520 turku, finland; 3laboratory for paleoenvironmental reconstruction, faculty of sciences, university of novi sad, trg dositeja obradovica 2, 21000 novi sad, serbia; 4recetox, faculty of science, masaryk university, kamenice 5, 62500 brno, czech republic *corresponding author: jussi.meriluoto@abo.fi no nco mm er cia l u se on ly t. kaloudis et al.2 mc/nod were associated with microcystis aeruginosa, planktothrix rubescens and nodularia spumigena strains. mc/nod were detected also in a m. wesenbergii strain, while p. agardhii strains and blooms were mostly negative for hepatotoxin production. first report of cyanobacterial paralytic shellfish toxin biosynthesis genes and paralytic shellfish toxin production in polish freshwater lakes by savela, spoof, höysniemi, vehniäinen, mankiewicz-boczek, jurczak, kokociński and meriluoto is a broad study on the cyanobacterial community of 34 lakes in western poland. it identified an established subpopulation of potential paralytic shellfish toxin (pst) producers, through the analysis of pst biosynthesis genes by pcr and qpcr, and pst production by hplc-fld in environmental samples. cyanobacterial dynamics and toxins concentrations in lake alto flumendosa, sardinia, italy by stefanelli, scardala, cabras, orrù, vichi, testai, funari and manganelli presents the long-term characterization of a cyanobacterial community and mc production in lake alto flumendosa (sardinia). the research demonstrated the predominance of p. rubescens, microcystis botrys and woronichinia naegeliana, and the significant persistence of toxic populations. cyanobacteria and cyanotoxins in drinking water – occurrence, monitoring, removal methods can cyanobacteria infect underground water sources? indications from small scale monitoring of a natural drinking water source (short note) by gkelis and vlamis presents the results of a small scale monitoring program in a greek water company demonstrating for the first time the presence of cyanobacteria in bottled natural mineral drinking water. the results indicate a potential hazard in this kind of product. cyanobacteria and microcystin contamination in untreated and treated drinking water in ghana by addico, hardege, kohoutek, degraft-johnson and babica shows frequent occurrence of cyanobacteria and mcs in nontreated water in treatment plants in ghana. water treatment significantly eliminated both cyanobacteria and mcs but toxins (maxima around 0.8 microgram per liter) were still found in approximately 15-20% samples of the treated water. chlorination and ozonation differentially reduced the microcystin content and tumour promoting activity of a complex cyanobacterial extract by sovadinová, babica, adamovský, alpatova, tarabara, upham and bláha demonstrated that ozone effectively removed all mcs from an extract of microcystis sp., and significantly reduced the overall tumour promotional potency of the sample. chlorination was much less effective and high doses of chlorine further produced toxic by-products. cyanotoxin detection techniques non-competitive elisa with broad specificity for microcystins and nodularins by akter, vehniäinen, meriluoto, spoof and lamminmäki reports the development of an easy-to-perform assay for generic detection of mcs and nod. the recombinant anti-immunocomplex antibody based non-competitive elisa was capable of detecting eleven toxin variants (mc-lr, -dmlr, -rr, -dmrr, -yr, la -ly, -lf -lw, -wr, and nod-r) below the who guideline concentration for mc-lr. cyanobacteria and cyanotoxins long term monitoring/reviews for specific geographical regions assessment of cyanoprokaryote blooms and of cyanotoxins in bulgaria in a 15-years period (2000-2015) by stoyneva-gäertner, descy, latli, uzunov, pavlova, bratanova, babica, maršálek, meriluoto and spoof presents a summary of results from studies carried out on 120 water bodies in bulgaria, where the cyanoprokaryote diversity was quite high (210 taxa of 60 genera). blooms were recorded in 14 and cyanotoxins were detected in 16 water bodies including 3 drinking water reservoirs. review of 130 years of research on cyanobacteria in aquatic ecosystems in serbia presented in a serbian cyanobacterial database by svirčev, tokodi and drobac presents an overview of the unique database concerning cyanobacterial distribution, cyanotoxin production and associated biological effects in different types of water bodies throughout the republic of serbia. review/synoptic paper of cyanocost research and activities toxic cyanobacteria and cyanotoxins in european waters – recent progress achieved through the cyanocost action and challenges for further research presented by 21 authors is a review summarising the outcomes of recent european research concerning toxic cyanobacteria and cyanotoxins, with an emphasis on developments within the framework of cyanocost. it highlights achievements and challenges for the phycological and ecological studies, analytical and detection approaches, toxicological research, management of toxic blooms as well as practices for cyanotoxin removal. no nco mm er cia l u se on ly foreword 3 acknowledgments this publication is based upon work from cost action cyanocost, supported by cost (european cooperation in science and technology). we would like to express our sincere thanks to all the themed issue authors. we knew the cyanocost community has potential for scientific work of the very highest quality but we were still positively surprised when we saw the great interest for this themed issue expressed by numerous authors across europe. we believe that many of the articles will have long-lasting usefulness for the cyanocost community as well as for a broad audience of international researchers and experts. we want to acknowledge and congratulate the four guest editors of this themed issue, pavel babica, camilla capelli, damjana drobac and spyros gkelis, for their highly professional editorial work. the co-editorin-chief of advances in oceanography and limnology nico salmaso is cordially thanked for his help and goodwill in realising the themed issue. our further thanks go to those many people who have had various responsibilities within the cyanocost action: the working group leaders and deputy leaders, those who produced handbooks and another themed issue within the action (see the review paper by meriluoto et al. in this themed issue). we want to express our gratitude to those who were responsible for the financial and administrative parts of the action, those who organised meetings and training schools, those who prepared and hosted the short-term scientific missions, and all others making cyanocost a success story. science officer dr. deniz karaca, administrative officer ms. tania gonzalez ovin and rapporteur mr. dick blaauboer from the cost organisation are warmly thanked for their support. triantafyllos kaloudis chairman of cyanocost jussi meriluoto working group 1 leader of cyanocost ludek blaha vice-chairman of cyanocost funded by the horizon 2020 framework programme of the european union no nco mm er cia l u se on ly layout 1 introduction cyanobacterial harmful algal blooms represent one of the most conspicuous waterborne microbial hazards to freshwater and marine ecosystems (codd et al., 2005; paerl et al., 2011). this hazard results from the production of cyanotoxins, harmful secondary metabolites, which can have deleterious effects within reservoirs and in downstream receiving water systems during releases (paerl and otten, 2013). harmful cyanobacterial blooms have increased globally in frequency and intensity in recent decades. eutrophication and warmer temperatures are often cited as key factors which promote these events (hudnell and dortch, 2008; paerl and huisman, 2008; gkelis et al., 2014). in greece, the warm mediterranean climate favors cyanobacterial blooms in eutrophic waters, which may start in spring and last until december; increased temperatures due to global warming may further enhance cyanobacteria dominance and promote toxic over non-toxic strains (gkelis et al., 2014). after the elucidation of cyanotoxins genes clusters, several studies have applied molecular methods for monitoring the presence of toxic cyanobacteria and the genes involved in the biosynthesis of cyanotoxins (hisbergues et al., 2003; vasconcelos et al., 2010; gkelis and zaoutsos, 2014). furthermore, molecular methods based on 16s rrna gene amplification are widely employed for the analysis of natural samples and/or monitoring freshwaters (kormas et al., 2011; loza et al., 2013). bottled water can come from a variety of sources including natural aquifers, springs, glacier run-off, and municipal water supplies, and these sources could potentially be contaminated with microcystins (cfia, 2011). to our knowledge, only two studies on levels of microcystins in bottled water have been published. those studies, performed in italy (ferretti et al., 2007) and canada (cfia, 2011; cfia, 2012), analysed domestic bottled water samples for the presence of microcystins and nodularin, and did not detect cyanotoxins. this work presents the results of a small scale monitoring program for drinking natural mineral water and highlights the necessity to initiate research and possibly establish official monitoring programs in cases where similar results are obtained in order to cope with the new demands for drinking water. methods sample collection and preparation the bottled natural mineral water originated from a water source (250 m depth), which is located in the prefecture of central macedonia, northern greece (n: 40°15’4.88” and e: 22°32’12.71”) with a water temperature ranging between 16°c in winter and spring and 18°c in autumn and summer. water samples were collected advances in oceanography and limnology, 2017; 8(1): 87-91 article doi: 10.4081/aiol.2017.6280 this work is licensed under a creative commons attribution-noncommercial 4.0 international license (cc by-nc 4.0). can cyanobacteria infect underground water sources? indications from small scale monitoring of a natural drinking water source spyros gkelis,1* aristidis vlamis1,2 1department of botany, school of biology, aristotle university of thessaloniki, gr-541 24 thessaloniki, greece; 2department of pharmacology, veterinary school, university of santiago de compostela, lugo 27002, spain *corresponding author: sgkelis@bio.auth.gr abstract the expansion of harmful cyanobacterial blooms is of worldwide concern as they have increased globally in frequency and intensity in recent decades. a cyanobacterial colony was found in a bottle of natural mineral water of a small water company in july 2012, which led to a further examination for a period of five months (july-november 2012) of both the bottled filtered water and the originating groundwater source (n. greece) for the occurrence of cyanobacteria. cyanobacteria occurrence was monitored by microscopy and cyanospecific 16s rdna amplification; potentially toxic species occurrence was screened by mcya gene (known to take part in the mc-biosynthetic gene cluster) amplification. the highest abundance of cyanobacterial cells without the simultaneous presence of the mcya gene, was measured in july, in contrast to october when the presence of cyanobacteria was only identified by tracing cyanospecific 16s rdna and the mcya gene region in the underground water source. the results of this small scale monitoring program indicate the potential existence of an emerging danger for human health in a relatively manageable product such as the bottled natural mineral water. key words: microcystis; microcystin; bottled water; natural mineral water. received: 12 september 2016. accepted: 16 march 2017. no nco mm er cia l u se on ly s. gkelis and a. vlamis88 monthly between july and december 2012 from two drilling points (sup1 and sup2) used for the production of bottled natural mineral water. the final product which had undergone filtration was also sampled (bottled filtered water-bf). aliquots of these were preserved with both lugol’s solution (1% v/v) and formaldehyde (2% v/v). water samples were stored in polyethylene bottles and transferred to the laboratory (<5 h) under cool and dark conditions. immediately upon reception in the laboratory, 1.5 l of water was filtered on a whatman gf/c filters and the filter was stored at -20°c for further analysis. phytoplankton analysis fresh and preserved samples were examined using an inverted microscope (olympus ix71) with phase-contrast. species were identified using komárek and anagnostidis (1999). the abundance of cyanobacterial cells was determined in accordance with utermöhl (1958), using 50 ml sedimentation chambers, thus detection limit was 20 cells l–1. transepts were counted and the variation coefficient was always kept under 20%. phytoplankton abundance is presented in number of cells l–1. molecular detection dna was extracted using the protocol described in atashpaz et al. (2010) for gram negative bacteria, after slicing the filters with a sterile scalpel. in order to identify cyanobacteria and potentially mc-producing cyanobacteria we used two different sets of primers, respectively: the 16s 27f (5’-agagtttgatcctggctcag-3’) /16s 1494r (5’tacggttaccttgttacgac -3’) primer pair which amplifies a 1367-bp fragment of 16s rdna in all cyanobacteria (neilan et al., 1997) and the mcya cd1f (5’aaaattaaaagccgtatcaaa-3’) /mcya cd1r (5’aaaagtgttttattagcggctcat-3’) primer pair (hisbergues et al., 2003), which was designed to amplify a 297-bp fragment of the mcya gene from mc-synthesizing cyanobacteria strains and was previously proved to be suitable to detect mc-producing cells from the genera anabaena, microcystis and, planktothrix (hisbergues et al., 2003). samples giving positive results in this assay have been shown to have a high probability of producing mcs (hisbergues et al., 2003; vasconcelos et al., 2010). we chose to detect a gene target known to be involved in the biosynthesis of mc, as this is the most frequent cyanotoxin in greece (gkelis and zaoutsos, 2014; gkelis et al., 2015b) and worldwide. pcr was carried out on the dna extracts using the primer pairs presented previously. all pcr reactions were prepared as described in gkelis and zaoutsos (2014). thermal cycling was carried out using an eppendorf mastercycler pro (eppendorf). amplification was performed according to the protocols described by neilan et al. (1997) and hisbergues et al. (2003) for 16s rrna and mcya, respectively. dna extracted from microcystis aeruginosa m6 strain (see vasconcelos et al., 2010) was used as positive control for the amplification of mcya gene target and water as negative control. pcr products were separated by 1.5% (w/v) agarose gel in 1x tae buffer. the gels were stained with ethidium bromide and photographed under uv transillumination. results and discussion cyanobacteria were detected by microscopy only in the bottled sample bf/a, collected on the 5th of july 2012 (tab. 1). a colonial form of microcystis-like cyanobacteria was found, which could not be further identified (fig. 1). cells were spherical with an average diameter tab. 1. microscopic and molecular detection of cyanobacteria in the water samples collected during the study. sampling date sample microscopic analysis μolecular analysis cells l–1 dna cyanobacteria 16s rrna mcya gene 05-07-12 sup (1) sup (2) bf/ a* 50,000 + + bf/ b + + 18-09-12 sup (1) sup (2) bf 31-10-12 sup (1) + + + bf + + 29-11-12 sup (1) sup (2) + bf + bf, bottled filtered water; sup (1), source underground water (old drill); sup (2), source underground water-(new drill); *sample bf/ a is from the same production line with bf/ b but sampled at a different time of the day. no nco mm er cia l u se on ly cyanobacteria in underground water sources 89 of 8.53μm (n=30, min 6.54 μm, max 11.6 μm). cyanobacterial abundance in this sample was 50,000 cells l–1, whereas in bf/b, collected on the same day and drilling source but at a different time, no cyanobacteria cells were found (tab. 1). no presence of cyanobacteria was observed by microscopy in any of the other samples, both bottled and from the underground water source. although the presence of algae and cyanobacteria in underground habitats (reisser, 2007) and the ability of soil and deep subsurface cyanobacteria to actively follow a moving water pocket in order to exploit it (garcia-pichel and pringault, 2001) has been shown, to the best of our knowledge this is the first report of planktonic cyanobacteria found in an underground drinking-water source. recently, pazouki et al. (2016) demonstrated that cyanobacteria can be found in water filtered through bank filtration, especially species with cell size <10 μm, such as the microcystis cells. in our case the source of the cyanobacterial colony we found remains unknown; however, it has been found that in the nearby river-reservoir system of aliakmon-polyphytos, microcystis aeruginosa can be dispersed in short distances through the wind (chrisostomou et al., 2009). furthermore, airborne cyanobacteria have been found in the nearby city of thessaloniki (genitsaris et al., 2011). nevertheless, since the source of the cyanobacterial colony was not found the possibility of opportunistic airborne contamination during the bottling cannot be excluded. dna was below the detection limit in seven out of the twelve samples analyzed (tab. 1). a pcr product of about 1370 bp was obtained using the 16s 27f/16s 1484r primer pair in four samples (including sample bf/a of 05-07-2012, where cyanobacteria were identified by microscopy). the 300 bp mcya-cd 1f/mcya-cd 1r primer pair pcr product, indicating the presence of mcya gene, was identified only in one out of the twelve samples assayed (tab. 1). in the studied water source the highest cyanobacterial abundance of 50.000 cells l–1 was observed in july without the simultaneous presence of the mcya gene responsible for mc production. at the end of october, however, the presence of cyanobacteria was only traced through the detection of 16s rdna as well as traces of the gene mcya. similar results were reported by davis et al. (2009) in lake agawam in july where the presence of toxic microcystis cells did not coincide with the presence of mc and also in lake champlain where mcs were detectable in october but without the presence of toxic microcystis cells at the same time. gkelis and zaoutsos (2014) found that the mcya region was amplified only where microcystis spp. were dominant and mc concentrations were >40 μg l–1. conjointly, gkelis et al. (2014) in a 14-month monitoring of lake pamvotis found that mcya was amplified only in two samples, where microcystis aeruginosa was dominant, whereas mcyb and mcye regions were amplified in almost all samples. several studies report that cyanobacteria can produce toxins at low temperatures but in combination with other important factors such as presence of light and nutrients’ availability (quiblier et al., 2013). the low temperatures (max. 18°c) of the studied groundwater source together with the absence of one of the genes responsible for cyanobacterial toxicity (mcya) in most of the samples suggest that the detected cyanobacteria do not probably constitute an important hazard for this specific source, also due to the limited input of nutrients and the absence of light. nevertheless, since traces of the mcya gene were detected in one sample, there is still a possibility that mcs could have been produced in the studied water source, since cyanotoxins are intracellular toxins contained within living cells (sivonen and jones, 1999) and cyanobacterial cells were found in high abundances in one sample. in this case, the biggest problem in bottled drinking natural mineral waters arises from the fact that the only treatment option in order to maintain the product type ‘natural mineral water’, is filtering, as filters can retain a big proportion of the microalgae but not the potential toxins produced by them resulting in human intoxication. our finding, although scarce, raises the question of monitoring algae in underground water sources. at present, none of the european countries have established monitoring program for cyanotoxins in potable minerals waters so far, and only some countries have done so for drinking water such as spain, france, the czech republic and poland (burch, 2008). in greece, there are very few official monitoring programs in place (kaloudis et al., 2013; gkelis et al., 2015a) for cyanobacterial blooms and toxins produced in freshwaters, whereas there is also no legislafig. 1. microphotograph of the microcystis-like colony found in the bottled filtered water bf/ a sample. no nco mm er cia l u se on ly s. gkelis and a. vlamis90 tion with regard to monitoring of these quality parameters (cook et al., 2005). moreover, there are no legally established maximum allowable concentrations for cyanotoxins in potable mineral water. the only relevant reference is in the national hygienic regulation a1β/4841/1979 (fek 696/β΄/1979), where the absence of microalgae is required for bottled waters intended for human consumption. a very small number of water treatment plants in greece control cyanobacterial growth, measure cyanotoxins or use water treatments for toxin removal: the athens water supply and sewerage company implements some measures for control and monitoring of cyanobacteria and cyanotoxins (kaloudis et al., 2013). other water utilities are small, at a local level, and they do not implement such measures, with the possible exception of the thessaloniki water supply and sewerage company (thessaloniki) (kaloudis, personal communication). conclusions the results of this small scale monitoring program demonstrated for the first time the presence of cyanobacteria in bottled natural mineral drinking water. while it seems unlikely that the use of bottled water would constitute any major hazard with regard to cyanotoxin exposure, our findings call for further research to investigate the presence, heterotrophic growth and significance of cyanobacterial colonies and/or biofilms in water distribution systems, such as wells. acknowledgments we thank prof. vitor vasconcelos for providing freezedried material of cyanobacteria strains used as positive control in pcr studies. av would like to thank dr. panagiota katikou for providing useful information on cyanotoxins. the authors acknowledge cyanocost-cost es 1105 for sharing of knowledge and networking and thank the two anonymous reviewers for helpful suggestions. references atashpaz s, khani s, barzegari a, barar j, vahed sz, azarbaijani r, omidi y, 2010. a robust universal method for extraction of genomic dna from bacterial species. microbiology 79:538-542. burch md, 2008. effective doses, guidelines and regulations. in: h.k. hudnell (ed.), cyanobacterial harmful algal blooms: state of the science and research needs. adv. exp. med. biol. 610:831-854. canadian food inspection agency (cfia), 2011. microcystins in bottled water. food safety action plan report. tschem-10/11, pp. 1-8 canadian food inspection agency (cfia), 2012. microcystins and nodularin in bottled water. food safety action plan report. ts-chem-11/12, pp. 1-8. chrisostomou a, moustaka-gouni m, sgardelis s, lanaras t, 2009. air-dispersed phytoplankton in a mediterranean river-reservoir system (aliakmon-polyphytos, greece). j. plankton res. 31:877-884. cook cm, moustaka-gouni m, gkelis s, lanaras t, 2005. greece: cyanotoxin risk assessment, risk management and regulation, p. 69-75. in: i. chorus (ed.), current approaches to cyanotoxin risk assessment, risk management and regulations in different countries. series wabolu 02/05. umweltbundesamt, dessau. codd ga, lindsay j, young fm, morrison lf, metcalf js, 2005. harmful cyanobacteria: from mass mortalities to management measures, p. 1-23. in: j. huisman, h.c.p. matthijs and p.m. visser (eds.), harmful cyanobacteria. springer, dordrecht. davis tw, berry dl, boyer gl, gobler cj, 2009. the effects of temperature and nutrients on the growth and dynamics of toxic and non-toxic strains of microcystis during cyanobacteria blooms. harmful algae 8:715-725. ferretti e, lucentini l, veschetti e, bonadonna l, stammati a, turco l, ottaviani m, 2007. screening and identification of unknown contaminants in water destined to human consumption: a case study. microchem. j. 85:57-64. garcia-pichel f, pringault o, 2001. cyanobacteria track water in desert soils. nature 412:380-381. genitsaris s, kormas ka, moustaka-gouni m, 2011. airborne algae and cyanobacteria: occurrence and related health effects. front. biosci. 3:772-787. gkelis s, chronis i, kagalou i, 2015a. monitoring a newly reborn patient: phytoplankton based water quality and cyanotoxin occurrence in a reconstructed shallow mediterranean lake. proceedings aslo aquatic sciences meet., granada. abstract id27076. gkelis s, lanaras t, sivonen k, 2015b. cyanobacterial toxic and bioactive peptides in freshwater bodies of greece: concentrations, occurrence patterns, and implications for human health. mar. drugs 13:6319-6335. gkelis s, papadimitriou t, zaoutsos n, leonardos i, 2014. anthropogenic and climate-induced change favors toxic cyanobacteria blooms: evidence from monitoring a highly eutrophic, urban mediterranean lake. harmful algae 39:322-333. gkelis s, zaoutsos n, 2014. cyanotoxin occurrence and potentially toxin producing cyanobacteria in freshwaters of greece: a multi-disciplinary approach. toxicon 78:1-9. hisbergues m, christiansen g, rouhiainen l, sivonen k, borner t, 2003. pcr-based identification of microcystinproducing genotypes of different cyanobacterial genera. arch. microbiol. 180:402-410. hudnell kh, dortch q, 2008. a synopsis of research needs identified at the interagency, international symposium on cyanobacterial harmful algal blooms (isoc-hab). adv. exp. med. biol. 619:17-43. kaloudis t, zervou ks, tsimeli k, triantis t, fotiou t, hiskia a, 2013. determination of microcystins and nodularin (cyanobacterial toxins) in water by lc–ms/ms. monitoring of lake marathonas, a water reservoir of athens, greece. j. hazard. mat. 263:105-115. komárek j, anagnostidis k, 1999. [cyanoprokaryota 1. teil: no nco mm er cia l u se on ly cyanobacteria in underground water sources 91 chroococcales], pp. 548. in: h. ettl, g. gärtner, h. heynig, and d. mollenhauer (eds.), [süsswasserflora von mitteleuropa 19/1].[book in german]. gustav fischer verlag, jena. kormas ka, gkelis s, vardaka e, moustaka-gouni m, 2011. morphological and molecular analysis of bloom-forming cyanobacteria in two eutrophic, shallow mediterranean lakes. limnologica 41:167-173. loza v, perona e, mateo p, 2013. molecular fingerprinting of cyanobacteria from river biofilms as a water quality monitoring tool. applied environ. microbiol. 79:1459-1472. neilan ba, jacobs d, deldot t, blackall ll, hawkins pr, cox pt, goodman ae, 1997. rrna sequences and evolutionary relationships among toxic and nontoxic cyanobacteria of the genus microcystis. int. j. syst. bacteriol. 47:693-697. paerl hw, hall ns, calandrino es, 2011. controlling harmful cyanobacterial blooms in a world experiencing anthropogenic and climatic-induced change. sci. total environ. 409:1739-1745. paerl hw, huisman j, 2008. blooms like it hot. science 320: 57-58. paerl hw, otten gt, 2013. harmful cyanobacterial blooms, causes, consequences, and controls. microb. ecol. 65: 995-1010. pazouki p, prévost m , mcquaid n, barbeau b, de boutray ml, zamyadi a, dorner s, 2016. breakthrough of cyanobacteria in bank filtration. water res. 102:170-179. quiblier c, wood s, echenique-subiabre i, heath m, villeneuve a, humbert j-f, 2013. a review of current knowledge on toxic benthic freshwater cyanobacteria-ecology, toxin production and risk management. water res. 47:5464-5479. reisser w, 2007. the hidden life of algae underground, p. 4758. in: j. seckbach (ed.), algae and cyanobacteria in extreme environments. springer, dordrecht. sivonen k, jones g, 1999. cyanobacterial toxins, p. 113-153. in: i. chorus and j. bartram (eds.), toxic cyanobacteria in water: a guide to their public health consequences, monitoring and management. world health organization, e & fn spon. utermöhl h, 1958. [zur vervollkommung der quantitativinen phytoplankton-methodik].[article in german]. int. ver. theor. angew. limnol. 9:1-38. vasconcelos v, martins a, vale m, antunes a, azevedo j, welker m, lopez o, montejano g, 2010. first report on the occurrence of microcystins in planktonic cyanobacteria from central mexico. toxicon 56:425-431. no nco mm er cia l u se on ly layout 1 “lightly we passed on earth” is the title of a novel by sergio atzeni, published in 1996, one year after his untimely death by drowning in the tyrrhenian sea. the author is referring to the levity of water. «lightly we passed on earth”…. “like the water that flows, jumps from the hollow spring, winds between mosses and ferns up to the roots of almond and cork trees or rolls over the stones from mountains and hills toward the plain, from stream to river, slowing down on the way to the swamps and the sea, transmuted by the sun in steam, in clouds moved by the wind, and in blissful rain…” giuseppe, peppe for family and friends, loved the water with which he felt in perfect harmony and in which he found comfort and peace also through the worsening of his illness. the attraction for water became scientific passion when peppe started his degree thesis under the supervision of delio ruggiu, a prominent italian limnologist. at the pallanza istituto idrobiologico, peppe found a place celebrated for its pioneering studies on lakes, their evolution, the importance of long term data, the significance and limitation of trophic indicators, the value of comparative limnology. to enter the pallanza istituto meant to immerse in the forefront of the international scientific debate on aquatic ecology having as colleagues of your own supervisor the reynolds, the harris, the vollenweider… in the silence broken only by the ticking of pierisa’s counter, to whom was confided the phytoplankton microscopic analyses and the field data compilation, lakes were investigated as ecosystems. in primis, lago maggiore where primary production measurements were pioneered in the 1950s by rich vollenweider during his stay at the istituto. then lago d’orta, so fragile, so needy of attention. lago d’orta was indeed peppe’s research topic for his degree thesis. at once, peppe demonstrated rare capacities of synthesis, critical thinking, associated with a gift for clear and effective writing. delio could only add few comments to peppe’s masterful presentation. in the second-floor laboratory, open to the majestic view of lago maggiore, student and supervisor built a scientific partnership that would produce a wealth of substantial contributions to phytoplankton ecology. in particular, peppe significantly contributed to the morpho-functional characterization of phytoplankton of subalpine deep lakes. this step was crucial for comparative limnology to understand structure and function of aquatic ecosystems under natural conditions, stress, and recovery. many colleagues have passed through this laboratory over the years to discuss of phytoplankton ecology: letizia garibaldi, nico salmaso, renata trevisan, luigi naselli flores, marco simona, fabio buzzi, giulio di leo; laurence carvalho, antonella lugliè, nicola sechi, just to cite a few. during her visits in italy, antonella cattaneo always visited the laboratory to discuss science and collaborate on studies of lago d’orta and lago di candia. in the context of an italy-quebec scientific cooperation program, peppe was invited by the biology department of université de montréal to collaborate with antonella and her colleague bernadette pinel-alloul in a comparative study of methods and biota of these two very different regions. peppe was a mentor to several students for their degree or phd theses. several of these students became in turn successful scientists and collaborators as alessandro oggioni with whom peppe shared the interest in the use of long-term data and historical series and the fine tuning of modern environmental technologies. moreover, peppe collaborated with martina austoni profiting of her expertise on phytoplankton taxonomy and photographic documentation of algal blooms and with tommaso sforzi through their effort to organize and archive phytoplankton monitoring data collected by environmental agencies in several italian lakes. alessandra pugnetti, that preceded peppe as a research staff of the cnr-istituto italiano di idrobiologia, shared with him many scientific discussions and research campaigns. with alessandra, peppe extended his research to the phytoplankton of high altitude alpine lakes, crucial habitats for studying the effects of acid rain and advances in oceanography and limnology, 2017; 8(2): 243-247 remembrance doi: 10.4081/aiol.2017.7209 this work is licensed under a creative commons attribution-noncommercial 4.0 international license (cc by-nc 4.0). in memoriam of peppe morabito firenze, april 16, 1964 castelletto ticino, july 12, 2017 no nco mm er cia l u se on ly m.m. manca244 contaminant long-distance transport on lake ecosystems. this research resulted in a seminal contribution to long term studies by implementing investigations started by rina monti stella in the 1880s. always with alessandra, peppe addressed the debate over the mechanisms at the base of the decoupling between primary production, i.e. potential growth, and real growth they highlighted that process variables responded better than standing variables, as chlorophyll and biomass, to nutrient reduction and oligotrophication, making worldwide famous the example of lago maggiore. when alessandra later moved to oceanography, their fruitful scientific partnership continued in the lter (long-term ecological research network) context with outstanding, some very recent, contributions to comparative studies in freshwater and marine ecosystems on development of advanced instrumentation and of theories on the role of climate and anthropogenic variables. peppe’s passion for field and laboratory studies and implementation of innovative technologies found a suitable application in a team research on lago di candia, a shallow lake in the canavese region of piedmont. the studies on this lake aimed at controlling the eutrophication through the integrated management of the lake watershed and its biota. lago di candia was thus offering the possibility to test the applicability of biomanipulation to lake remediation while assessing the relative importance of biotic and abiotic variables and the validity of the size efficiency hypothesis of brooks & dodson. working on this case study, that became the topic of his phd thesis at the università di parma, peppe had the opportunity to express his originality, combining classic phytoplankton methods with cutting-edge laboratory tools and in situ incubations. he also showed a clear and singularly elegant writing style in the presentation and discussion of the findings. the thesis earned him, infra alia, a prize that he received thanking with his trademark modesty his teachers (among whom ireneo ferrari) and fellow students for their contribution to his research. for me, lago di candia remains linked to a personal souvenir of peppe's generosity and kindness. in order to study zooplankton diel vertical migration (dvm), we had planned an intensive research effort with samplings scheduled every 4 h, day and night. because each sampling lasted 2 h, only 2 h remained between successive outings. we were three researchers: peppe, waleed hamza (a phd student), and myself. waleed and i collected the zooplankton in the boat while peppe was processing the samples in the base-camp laboratory. at 2 am, we had concluded a sampling and the next one was planned for 4 pm. we were sleeping in sleeping bags under an awning close to the lake and i was suffering of a convulsive gravidic cough. fearing of being overcome by exhaustion, i asked my colleagues to my colleagues to rouse me, if still asleep. in the morning, i woke up to realize that the 4 am sampling had been completed by peppe and walid while i was fast asleep. i will never forget the brightness of peppe’s smile when i thanked him. alessandra, peppe and waleed went to a congress on dmv in amsterdam when my son dario was just one month old. they brought me back a wonderful poster with the cladoceran drawings from sars’ book. that poster with its touching inscription is still with me. the multi-faceted expertise of peppe on phytoplankton (relationships between structure and function, development of algal indicator species, succession, long-term trends under trophic and climatic stresses, ecophysiology) was precious in the implementation of the water framework directive, in which the quality of freshwater bodies is based on ecological criteria, often centered on phytoplankton. this approach, carried out from 2004 to 2013, brought peppe to oversee several large scale european projects as “local hydro-morphology, habitat and river basin management plans (rmbs): new measures to improve ecological quality in south european rivers his and lakes (life+ inhabit); water bodies in europe: integrative systems to assess ecological status and recovery (eu fp-7: wiser); “evaluating the impacts of global change on european freshwater ecosystems (eu fp-6: euro-limpacs). scientific interest was combined in peppe with a strong civic commitment. he believed that being a scientist should not be limited to interaction with colleagues in national and international meetings and discussion groups but should also entail the responsibility of educating the public to the problems faced by inland waters. in a slide prepared for one of his several educational presentations, he warned: “lakes represent the largest reserve of immediately available freshwater on the planet; this reserve is sufficient but not abundant…”. furthermore, he wrote in the announcement of the aiol (associazione italiana di oceanologia e limnologia) congress organized in pallanza: “to foster the transfer of scientific findings to society, administrators at the local, regional, and national level should be invited to attend the congress. moreover, a section of the congress should be open to the public at large (especially the schools) to encourage interaction with the scientists to promote and popularize the knowledge about protection and management of water resources”. societal commitment, profound expertise in phytoplankton spatial and temporal dynamics, and appeal for modern oceanographic technologies led peppe to a novel no nco mm er cia l u se on ly in memory of peppe morabito 245 research, he dreamed of realizing in lakes, starting from maggiore and orta, the experience of ships of opportunity, collecting in real time relevant environmental data by means of sensors deployed on moving vessels. the project sailing (sensor based assessment of inlake processes and water quality: scientific innvestigation and growing environmental awareness) in his conception, realization and data analysis was a collaboration of peppe with michela rogora and dario manca, both at cnr ise, and with walter zerla, promoter of the initiative “sailing on lago maggiore and lago d’orta”. the connection between peppe, zerla and the other colleagues shines through in the photos taken during the project realization the interactions between the impact of global climatic factors and local anthropogenic disturbances has been pursued recently in the project blasco (blending laboratory and satellite techniques for detecting cyanobacteria in lakes). cyanobacterial blooms are increasingly observed worldwide generating preoccupation for lake trophic function and for safety of water use. this project has been considered outstanding by a panel of international referees “it is an excellent project based on a long tradition of limnological studies brought to a global level through satellite data. the findings would appeal a large community of limnologists. this research is timely because it addresses an urgent problem and promises to improve the calibration of satellite-based technologies”. i had the chance, as the director of the istituto, to follow the different stages of this project going from the inception to its development in which peppe collaborated with alessandro oggioni and mariano bresciani, both at irea cnr. this research linking field methods with sophisticated laboratory techniques has already resulted in significant international publications. blasco was peppe’s last project to which he dedicated all his time despite his worsening health. he succeeded to accomplish this research until the final workshop of june 15 2017, just twenty-seven days before his untimely death. this final achievement is one of his many scientific and humane legacies. i like to remember over all his particular way to be a man of science and a concerned citizen, to be profound in his research and light, never presumptuous, in the delivery of his results. these are rare virtues indeed, particularly nowadays. marina m. manca senior research associate, cnr ise verbania pallanza italy translated by antonella cattaneo universitè de montreal, montreal quebec, canada some selected publications bettinetti r, morabito g, provini a 2000. phytoplankton assemblage structure and dynamics as indicator of the recent trophic and biological evolution of the western basin of lake como (n. italy). hydrobiologia 435:177-190. boggero a, fontaneto d, morabito g, volta p, 2014. limnology in the 21st century: the importance of freshwater ecosystems as model systems in ecology and evolution. j. limnol. 73:1-3. bresciani m, giardino c, lauceri r, matta e, cazzaniga i, pinardi m, lami a, austoni m, viaggiu e, congestri r, morabito g, 2016. earth observation for monitoring and mapping of cyanobacteria blooms. case studies on five italian lakes. j. limnol. 76(s1):127-139. bresciani m, rossini m, morabito g, matta e, pinardi m, cogliati s, julitta t, colombo r, braga f giardino c, 2013. analysis of within-and between-day chlorophyll-a dynamics in mantua superior lake, with a continuous spectroradiometric measurement. mar. freshwat. res. 64:303-316. bresciani m, stroppiana d, odermatt d, morabito g, giardino c, 2011. assessing remotely sensed chlorophyll-a for the implementation of the water framework directive in european perialpine lakes. sci. total environ. 409:3083-3091. callieri c, caravati e, morabito g, oggioni a, 2005. the unicellular freshwater cyanobacterium synechococcus and mixotrophic flagellates: evidence for a functional association in an oligotrophic, subalpine lake. freshwater biol. 51:263-273. callieri c, morabito g, huot y, neale pj, litchman e 2001. photosynthetic response of picoand nanoplanktonic algae to uvb, uva and par in a high mountain lake. aquat. sci. 63:286-293. carvalho l, poikane s, lyche solheim a, phillips g, borics g, catalan j, de hoyos c, drakare s, dudley bj, järvinen m, laplace-treyture c, maileht k, mcdonald c, mischke u, moe j, morabito g, nõges p, nõges t, ott i, pasztaleniec a, skjelbred b, thackeray sj, 2012. strength and uncertainty of phytoplankton metrics for assessing eutrophication impacts in lakes. hydrobiologia 704:1-14. cattaneo a, de sève m, morabito g, mosello r, tartari g, 2011. periphyton changes over 20 years of chemical recovery of lake orta, italy: differential response to perturbation of littoral and pelagic communities. j. limnol. 70:177-185. gallina n, salmaso n, morabito g, beniston m, 2013. phytoplankton configuration in six deep lakes in the peri-alpine region: are the key drivers related to eutrophication and climate? aquat. ecol. 47:177-193. giardino c, bresciani m, stroppiana d, oggioni a, morabito g, 2014. optical remote sensing of lakes: an overview on lake maggiore. j. limnol. 73:210-214. lami, a, tartari ga, musazzi s, guilizzoni p, marchetto a, manca m, boggero a, nocentini am, morabito g, tartari g, guzzella l, bertoni r, callieri c, 2007. 21 high altitude lakes: limnology and paleolimnology. dev. earth surf. process. 10:155-170. lauceri a, bresciani m, lami a, morabito g, 2017. chlorophyll a interference in phycocyanin and allophycocyanin spectrophotometric quantification. j. limnol. e-pub 28 september 2017. doi: 10.4081/jlimnol.2017.1691 manca m, cavicchioni n, morabito g 2000. first observations no nco mm er cia l u se on ly m.m. manca246 on the effect of a complete, exceptional overturn of lake maggiore on plankton and primary productivity. int. rev. gesamten hydrobiol. hydrograp. 85:209-222. morabito g, 2001. six years’ (1992-1997) evolution of phytoplankton communities after recovery by liming in lake orta, northern italy. lakes reserv. res. manag. 6:305-312. morabito g, boggero a, galanti g, giussani g, oggioni a, volta p, 2008. problemi e prospettive per l’utilizzo delle biocenosi lacustri come indicatori di qualità ecologica ai sensi della direttiva comunitaria 2000/60/ce. progetto c 51. rapporto finale: 99 pp. morabito g, curradi m 1997. phytoplankton community structure of a deep subalpine italian lake (lake orta, n. italy) four years after the recovery from acidification by liming. int. rev. gesamten hydrobiol. hydrograp. 82:487-506. morabito g, oggioni a, austoni m, 2012. resource ratio and human impact: how diatom assemblages in lake maggiore responded to oligotrophication and climatic variability. hydrobiologia 698:47-60. morabito g, oggioni a, caravati e, 2005. decadal trends of pelagic algal biomass capacities in lago maggiore (n. italy). verh. internat. verein. limnol. 29:231-234. morabito g, oggioni a, caravati e, panzani p, 2007. seasonal morphological plasticity of phytoplankton in lago maggiore (n. italy). hydrobiologia 578:47-57. morabito g, oggioni a, panzani p, 2003. phytoplankton assemblage at equilibrium in large and deep subalpine lakes: a case study from lago maggiore (n. italy). hydrobiologia 502:37-48. morabito g, mazzocchi mg, salmaso n, zingone a bergami c, flaim g, accoroni s, basset a, bastianini m, belmonte g, bernardi aubry f, bertani i, bresciani m, buzzi f, cabrini m, camatti e, caroppo c, cataletto b, castellano m, -del negro p, de olazabal a, di capua i, elia ac, fornasaro d, giallain m, grilli f, leoni b, lipizer m, longobardi l, ludovisi a, lugliè a, manca m, margiotta f, mariani ma, marini m, marzocchi m, obertegger u, oggioni a, padedda bm, pansera m, piscia r, povero p, pulina s, romagnoli t, rosati i, rossetti g, rubino f, sarno d, satta ct, sechi n, stanca e, tirelli v, totti c, pugnetti a, 2017. plankton dynamics across the freshwater, transitional and marine research sites of the lter-italy network. patterns, fluctuations, drivers. sci.total environ. (in press). morabito g, pugnetti a 2000. primary productivity and related variables in the course of the trophic evolution of lake maggiore. int. ver. theor. angewan. limnol. verh. 27:2934-2937. morabito g, ruggiu d, panzani p 2001. trends of phytoplankton characteristics and their communities in preand postliming time in lake orta (1984-1998). j. limnol. 60:91-100. morabito g, ruggiu d, panzani p, 2002. recent dynamics (1995-1999) of the phytoplankton assemblages in lago maggiore as a basic tool for defining association patterns in the italian deep lakes. j. limnol. 61:129-145. morabito g, salmaso n, ruggiu d, 2003. phytoplankton association patterns in the deep southern subalpine lakes (part 2). j. limnol. 62:75-176. neale pj, litchman e, sobrino c, callieri c, morabito g, montecino v, huot y, bossard p, lehmann c, steiner d 2001. quantifying the response of phytoplankton photosynthesis to ultraviolet radiation: biological weighting functions versus in situ measurements in two swiss lakes. aquat. sci. 63:265-285. pahissa j, catalan j, morabito g, dörflinger g, ferreira j, laplace-treyture c, gîrbea r, marchetto a, polykarpou p, de hoyos c, 2015. benefits and limitations of an intercalibration of phytoplankton assessment methods based on the mediterranean gig reservoir experience. sci. total environ. 538:169-179. pareeth s, bresciani m, buzzi f, leoni b, lepori f, ludovisi a, morabito g, adrian r, neteler m, salmaso n, 2017. warming trends of perialpine lakes from homogenised time series of historical satellite and in-situ data. sci. total environ. 578:417-426. phillips g, lyche-solheim a, skjelbred b, mischke u, drakare s, free g, de hoyos mjc, morabito g, poikane s, carvalho l, 2012. a phytoplankton trophic index to assess the status of lakes for the water framework directive. hydrobiologia 704: 75-85. pugnetti a, camatti e, mangoni o, morabito g, oggioni a, saggiomo v, 2006. phytoplankton production in italian freshwater and marine ecosystems: state of the art and perspectives. chem. ecol. 22(s1):49-69. ruggiu d, morabito g, panzani p, pugnetti a 1998. trends and relations among basic phytoplankton characteristics in the course of the long-term oligotrophication of lake maggiore (italy). hydrobiologia 370: 243-257. salmaso n, buzzi f, cerasino l, garibaldi l, leoni b, morabito g, rogora m, simona m, 2013. influence of atmospheric modes of variability on the limnological characteristics of large lakes south of the alps: a new emerging paradigm. hydrobiologia 731:31-48. salmaso n, morabito g, buzzi f, garibaldi l, simona m, mosello r, 2006. phytoplankton as an indicator of the water quality of the deep lakes south of the alps. hydrobiologia 563:167-187. salmaso n, morabito g, garibaldi l, mosello r, 2007. trophic development of the deep lakes south of the alps: a comparative analysis. fundam. appl. limnol. 170:177-196. salmaso n, morabito g, mosello r, garibaldi l, simona m, buzzi f, ruggiu d, 2003. a synoptic study of phytoplankton in the deep lakes south of the alps (lakes garda, iseo, como, lugano and maggiore). j. limnol. 62:207-227. stefani f, salerno f, copetti d, rabuffetti d, guidetti l, torri g, naggi a, iacomini m, morabito g, guzzella l, 2016. endogenous origin of foams in lakes: a long-term analysis for lake maggiore (northern italy). hydrobiologia 767:249-265. tapolczai k, anneville o, padisák j, salmaso n, morabito g, zohary t, tadonléké rd, rimet f, 2014. occurrence and mass development of mougeotia spp. (zygnemataceae) in large, deep lakes. hydrobiologia 745:17-29. taranu ze, gregory-eaves i, leavitt pr, bunting l, buchaca t, catalan j, domaizon i, guilizzoni p, lami a, mcgowan s, moorhouse h, morabito g, pick fr, stevenson ma, thompson pl, vinebrooke rd, 2015. acceleration of cyanobacterial dominance in north temperate-subarctic lakes during the anthropocene. ecol. lett. 18:375-384. tessier c, cattaneo a, pinel-alloul b, galanti g, morabito g, 2004. biomass, composition and size structure of invertebrate communities associated to different types of aquatic vegetation during summer in lago di candia (italy). 63:190-198. no nco mm er cia l u se on ly in memory of peppe morabito 247 thackeray sj, nõges p, dunbar mj, birger skjelbred bjd, morabito g, carvalho l, phillips g, mischke u, catalan j, de hoyos c, laplace c, austoni m, padedda bm, maileht k, pasztaleniec a, järvinen m, lyche solheim a, clarke rt, 2013. quantifying uncertainties in biologically-based water quality assessment: a pan-european analysis of lake phytoplankton community metrics. ecol. indic. 29:34-47. tiberti r, metta s, austoni m, callieri c, morabito g, marchetto a, rogora m, tartari ga, von hardenberg j, provenzale a, 2013. ecological dynamics of two remote alpine lakes during ice-free season. j. limnol. 72:401-416. tolotti m, forsström l, morabito g, thaler b, stoyneva m, cantonati m, šiško m, lotter af, 2009. biogeographical characterisation of phytoplankton assemblages in high altitude, and high latitude european lakes. adv. limnol. 62:55-75. tolotti m, manca m, angeli n, morabito g, thaler b, rott e, stuchlik e 2006. phytoplankton and zooplankton associations in a set of alpine high altitude lakes: geographic distribution and ecology. hydrobiologia 562:99-122. trevisan r, picarella m, dazzo fb, bona s, morabito g, squartini a, 2014. using a morpho-functional approach to assess phytoplankton dynamics in two adjacent high-mountain lakes: a 10-year survey. j. limnol. 73:409-420. wolfram g, argillier c, de bortoli j, buzzi f, dalmiglio a, dokulil mt, hoehn e, marchetto a, martinez p-j, morabito g, reichmann ms, remec-rekar š, riedmüller u, rioury c, schaumburg j, schulz l, urbanič g, 2009. reference conditions and wfd compliant class boundaries for phytoplankton biomass and chlorophyll-a in alpine lakes. hydrobiologia 33:45-58. more at https://scholar.google.it/citations?user=qzxlu0maaaaj&hl=it no nco mm er cia l u se on ly layout 1 introduction prior to the proliferation of large dams in the second half of the 20th century, global rivers on average delivered about 19 billion tons (bt) of sediment annually to the ocean (milliman and farnsworth, 2011), the volumetric equivalent of about 15 or 16 great walls of china entering the ocean each year. one might assume that large rivers would be the dominant source for these sediments. the amazon and ganges-brahmaputra (1.2 bt yr–1) and yangtze and mississippi rivers (0.4-0.5 bt yr–1), for example, prior to river damming, were responsible for about 15% of the total global sediment flux. however, there is an increasing realization that smaller mountainous rivers discharge far more sediment than their drainage basin areas would suggest (milliman and syvitski, 1992; syvitski and milliman, 2007). if one normalizes the annual sediment load by dividing it by the river’s basin area (sediment yield), the amazon river (basin area 6,300,000 km2) and the pre-dam mississippi (basin area 3,300,000 km2) had annual sediment yields of ~190 t km–2 yr–1 and ~130 t km–2 yr–1, respectively. other examples can be found of large rivers characterized by even smaller sediment yields, such as the parana-paraguay system (argentina), where a basin spanning 2.6×106 km2 surface area has an annual sediment yield of ~50 t km–2 yr–1 (guerrero et al., 2015). in contrast to these large rivers, 14 small rivers on the island of java, whose combined basin areas total 52,000 km2, discharge about 100 million tons of sediment annually (milliman et al., 1999), which equates to an average sediment yield of 2000 t km–2 yr–1. even more impressive are 22 taiwanese rivers (drainage basin areas totaling 25,000 km2) that collectively discharge 200 mt yr–1 (kao and milliman, 2008; milliman and farnsworth, 2011, p. 307), equating to a sediment yield of 8000 t km–2 yr–1, 40fold greater than that of the amazon. globally there are 18 rivers with basin areas greater than 1,000,000 km2, compared to an estimated 4000 rivers with basin areas between 1000 and 10,000 km2. given their sheer number plus their often high sediment yields, small rivers collectively can have a disproportionately great effect on sediment flux to the global ocean, an effect that often has been underappreciated. small mountainous rivers, that is rivers with basin areas less than 10,000 km2 and draining mountains higher than 1000 m (milliman and syvitski, 1992), account for about 10% of the land area draining into the global ocean, but collectively they account for about 15% of the annual water discharge and 45% and 30% of the suspended and dissolved solids reaching the ocean annually (table 2.13 in milliman and farnsworth, 2011). small mountainous rivers are also much more likely to have suspended sediment concentrations at hyperpycadvances in oceanography and limnology, 2016; 7(2): 106-114 article doi: 10.4081/aiol.2016.5899 this work is licensed under a creative commons attribution-noncommercial 4.0 international license (cc by-nc 4.0). flux and fate of river-discharged sediments to the adriatic sea john d. milliman,1,2 davide bonaldo,2* sandro carniel2 1school of marine science, college of william and mary, 1375 greate road, gloucester point 23062-1346, va, usa; 2institute of marine sciences, national research council of italy, castello 2737/f, 30122 venice, italy *corresponding author: davide.bonaldo@ve.ismar.cnr.it abstract small rivers, particularly those draining mountainous terrain, discharge disproportionately large quantities of sediment to the global ocean. because small mountainous rivers are more susceptible to catastrophic events, they tend to discharge their sediments over relatively short periods of time, such as during floods. the impact of small mountainous rivers is especially evident on the coastal ocean, such as the adriatic sea where fully 75% of the estimated 145 million tons (mt) of discharged sediment comes from rivers with basin areas smaller than 7000 km2. within this semi-enclosed basin in the northeast of the mediterranean sea, of particular note are the high sediment loads of five albanian rivers (located in the southeast), which, prior to dam construction, collectively discharged about 85 mt yr–1 perhaps much of it at hyperpycnal concentrations, which would have allowed the sediment to bypass the shelf and be deposited at greater depths. geochemical data confirm that albanian river sediment extends well into the southern and central adriatic sea. delineating and understanding the flux and fate of adriatic sea sediments may be best facilitated through the reanalysis of existing river datasets and the acquisition of new river data, particularly during periodic floods, high-resolution seismic profiling coupled with sitespecific coring, as well as application of integrated ocean-sediment numerical models. key words: adriatic sea; river discharge; sediments; river sediment fate; albania. received: march 2016. accepted: june 2016. river sediment flux and fate in the adriatic sea 107 nal levels, that is >40 g l–1. when hyperpycnal river waters enter the ocean they are sufficiently dense to sink and carry the sediment across the shelf and into deeper water (mulder and syvitski, 1995). drake et al. (1972), for instance, found turbidite deposits in the adjacent santa barbara basin that had been deposited only a few weeks after the 1969 flood of the santa clara river mentioned above; similarly, mulder et al. (1998, 2003) found turbidite deposits seaward of the var river in southern france following a major flood. there are several reasons why small mountainous rivers have high sediment yields. with elevations higher than 1000 m and with steep gradients, together with relatively small flood plains, small-river basins experience high rates of erosion but have relatively little accommodation space in which the eroded sediment can be stored. moreover, small-river basins are more responsive to episodic events, such as earthquakes or floods. large river basins such as the amazon or mississippi can modulate a local storm or flood; the maximum measured discharge of the amazon, for instance, is only about twice its average discharge (fig. 3.45 in milliman and farnsworth, 2011). heavy rainfalls over small watersheds, in contrast, often can generate short-lived and flashy events. a typhoon-generated flood on the choshui river (taiwan) in july 2001, for instance, lasted only a few hours, but its peak discharge was 28,000 m3 sec–1, about 3 orders of magnitude greater than the river’s flow before or after the flood. such short-lived flashy floods, in fact, are responsible for more than 70% of the sediment discharge in most taiwanese rivers (kao and milliman, 2008). such events do not only occur in typhoon-prone areas, however. during several heavy rainstorms in the winter of 1969, for instance, the santa clara river, a small (4100 km2 basin area) semi-arid river about 80 km north of los angeles, discharged 45 mt of sediment, 18 mt on one day (fig. 1); most of the discharge occurred at hyperpycnal concentrations, peak concentrations exceeding 160 g l–1 (waananen, 1969; warrick and milliman, 2003). over the preceding 20 years, when rainfall was sparse and there were no floods, the santa clara discharged only about 15 mt in total. in contrast to the 18 mt discharged by the santa clara on a single day, during a peak flood year in 1950 the mississippi river (3,300,000 km2 basin area) discharged more than 500 mt of sediment, but daily sediment loads never exceeded 5 mt (fig. 1); moreover, suspended sediment concentrations never reached 1 g l–1. within the mediterranean basin, the adriatic sea represents a paradigmatic test site for investigating the role of small mountain rivers in the sediment budget of a relatively small, semi-enclosed basin, and their contribution to mass and nutrient fluxes off the continental shelf. with reference to this site, the present paper aims at providing some preliminary estimates for this process, assessing the main sources of uncertainty and suggesting possible strategies for filling this information gap. sediment discharge to the adriatic sea the adriatic sea is a nw-se-trending semi-enclosed basin within the northern mediterranean sea, approximately 200 km in width and 800 km in length (fig. 2). its northernmost area is constituted by a gently sloping continental shallow shelf, depths on the order of a few tens of meters, which is occasionally exposed to intense cold winds, spanning from the gulf of venice to ancona (fig. 2). the broad, shallow northern sub-basin is interrupted south of ancona by the pomo pit, a depression crossing the basin along its sw-ne axis that reaches bottom depths from 150 to 280 m. to the southeast off the apulian and albanian coasts, the shelf dips into the south adriatic pit (sap), the deepest region in the adriatic sea at 1200 m. the sap is connected by the palagruža sill (180 m) by the middle adriatic pit to the north, while the limit between the adriatic sea proper and the deeper ionian sea is defined by the otranto strait sill (780 m) to the fig. 1. daily sediment discharges from the santa clara (blue) and the mississippi (red) rivers during their record discharge years, 1969 and 1950, respectively. two flash floods, each lasting a total about 3-4 days, in the winter of 1969 accounted for 98% (45 mt) of the santa clara river’s annual sediment load, nearly all at hyperpycnal concentrations (>40 g/l). peak daily discharge was 18 mt; daily discharge for the remaining 358 days ranged from 0 (when the river was dry) to ~200,000 tons. the mississippi river in 1950, by contrast, discharged about 500 mt, daily discharge averaging more than 1 mt but never exceeding 4.5 mt; sediment concentrations never reached 1 g/l. reprinted from: j.d. milliman and k.l. farnsworth kl, 2011. river discharge to the coastal ocean: a global synthesis. cambridge university press; this figure cannot be reproduced, shared, altered, or exploited commercially in any way without the permission of cambridge university press, as it is copyrighted material and therefore not subject to the allowances permitted by an open access licence. 108 j.d. milliman et al. south. being an interface between the mediterranean and the adriatic basins, the southern adriatic margin (sam) acts as a cold engine for regional circulation, and thus serves as a crossroads for a variety of hydrodynamic, geological and biological processes (carniel et al., 2016a). few coastal seas in the world are as dominated by as many small mountainous rivers as the adriatic sea: alpine rivers (the po, adige, brenta, piave and tagliamento rivers) to the northwest, apennine rivers to the west, and albania mountainous rivers in the southeast. in total we count 35 rivers draining into the adriatic basin (fig. 3), for which we have suspended sediment data for 29 rivers (see p. 241-247 in milliman and farnsworth, 2011). the po river is by far the largest adriatic river, draining 74,000 km2 and having an average flow of 1500 m3 s–1; if one includes the drini (albania; 20,000 km2) and the adige (northeast italy; 12,000 km2), collectively these three rivers drain ~60% of the 173,000 km2 drained by the 35 adriatic rivers. headwaters in 11 of these rivers – notably those draining italian alps plus the albanian rivers lie at elevations >2000 m, the po being by far the highest at 4800 m. in contrast, 19 of the 35 rivers have basin areas less than 1500 km2, and 11 have headwaters lower than 1000 m in elevation; 8 of those rivers drain italy’s apennines. viewing the 35 adriatic rivers in total, sediment discharge from italian rivers shows a positive log-linear relationship with basin area, there being apparently little difference in sediment yield between lower and higher elevation rivers (fig. 4; blue diamonds). albanian rivers, in contrast, have much higher sediment loads relative to basin area. the shkumbini river (2200 km2) , for instance, has a sediment load of 7.2 mt yr–1, whereas the apennine ofanto river (2700 km2 basin area) has an average sediment load of 0.9 mt yr–1. more impressive are the 30 mt yr–1 discharged by the albanian semani river (5600 km2); in comparison, the adige in northern italy (basin area 12,000 km2) discharges only 1.6 mt yr–1. the high albanian sediment loads and yields, we assume, reflect heavy rains and resulting flashy floods, frequent earthquakes and land use. trincardi et al. (1994) and syvitski and kettner (2007), as well as the database in milliman and farnsworth, 2011 (see table c8), have estimated that prior to dam construction the northern italian rivers collectively discharged 34-45 million tons of sediment annually (mt yr–1) to the adriatic. considering the five alpine rivers alone (fig. 3), we estimate that prior to dam construction they discharged about 34 mt yr–1 (fig. 4). in addition, apennine rivers discharge another 22 mt yr–1, leading to a total of about 56 mt yr–1 of sediment discharged to the western side of the adriatic sea prior to dams (milliman and farnsworth, 2011). of greater sedimentological importance are five main albanian rivers, four of which have headwaters at elevations higher than 2000 m. collectively, prior to dam construction, these five rivers discharged an estimated 85 mt yr–1, the vijose and semani rivers each discharging more than 25 mt yr–1 (simeoni et al., 1997; ciavola et al., 1999) (fig. 5). based on available data, mulder and syvitski (1995) listed albania’s drini river as a potentially hyperpycnal river; but with further data, we now can include the pre-dammed semani and vijose rivers (sediment concentrations 10 and 5 mg l–1, respectively; see p. 247 in milliman and farnsworth, 2011) as capable of discharging suspended sediment at hyperpycnal levels. with the exception of the neretva river (14 mt yr–1; eurosion, 2004), croatian rivers, which in large part drain limestone terrain, discharge relatively little sediment to the adriatic. prior to dam construction, italian, croatian and albanian rivers therefore discharged about 145 mt yr–1 of sediment to the adriatic sea. in terms of basin size, the three large rivers (po, adige and drini) would have accounted for only 25% of the sediment discharged to the adriatic (15, 12 and 16 mt yr–1, respectively). the remaining 75% of the adriatic’s sediment input came from 32 rivers, their drainage basins all being less than 7000 km2 in area. albanian rivers contributed about 60% of the total sediment entering the adriatic (fig. 3); this flux, however, has declined with the construction of dams (simeoni et al., 1997; milliman and farnsworth, 2011) and probably will continue to decline with the future construction of more dams (bethge, 2014). one can therefore view the adriatic is having abundant sediment discharge to the southeastern and northwestern parts of the basin compared to relatively little discharge to the central sections of the basin. fig. 2. adriatic sea bathymetry and main sites discussed in the text. river sediment flux and fate in the adriatic sea 109 problematic sediment discharge data the sediment loads presented in figs. 4 and 5 are, at best, estimates of values prior to the dam construction that occurred largely in the latter decades of the 20th century. with few exceptions, such as the po river, the database is difficult to access, suggesting that, among other things, the data for many, perhaps most, rivers may be sparse in terms of valid observations and/or number of years of measurement. this problem is particularly critical for rivers in which annual discharge varies widely, such as those rivers subject to periodic flash floods. sudden, short-lived flash floods can raise the river flow an order of magnitude or more within a few hours – and decline almost as quickly (warrick and milliman, 2003; milliman and kao, 2005). being a non-linear function of water discharge, suspended sediment concentration and thus sediment discharge can increase by one to three orders of magnitude. monitoring flash floods, however, is extremely difficult, and only a few have ever been adequately monitored. often sediment rating curves during such events may have been based on one or two sediment samples which are then compared to daily water discharge, and using the resulting rating curve, daily sediment discharge could be computed (inman and jenkins, 1999). the problem with this approach is that both peak sediment concentration and water discharge can vary greatly hour-to-hour and by an order of magnitude or more within 24 h. fig. 3. adriatic sea and the 35 discharging rivers for which there are published data. reprinted from: j.d. milliman and k.l. farnsworth kl, 2011. river discharge to the coastal ocean: a global synthesis. cambridge university press; this figure cannot be reproduced, shared, altered, or exploited commercially in any way without the permission of cambridge university press, as it is copyrighted material and therefore not subject to the allowances permitted by an open access licence. 110 j.d. milliman et al. fig. 4. mean annual suspended sediment loads vs basin areas for 29 rivers that discharge into the adriatic sea; we have found no sediment discharge data for the other 6 adriatic rivers. upland and mountain (mtn) rivers have basins with 500-1000 m and 1000-3000 m average height respectively. data compiled by: j.d. milliman and k.l. farnsworth kl, 2011. river discharge to the coastal ocean: a global synthesis. cambridge university press; this figure cannot be reproduced, shared, altered, or exploited commercially in any way without the permission of cambridge university press, as it is copyrighted material and therefore not subject to the allowances permitted by an open access licence. fig. 5. mean annual (pre-dam) sediment flux to the adriatic sea from rivers draining the italian alps (5 rivers), italian apennines (24 rivers), croatia (1 river) and albania (5 rivers). sediment loads are in millions of tons per year; arrows are proportional to the loads. river sediment flux and fate in the adriatic sea 111 comparing daily and hourly water discharge data – rather than hourly discharge data in calculating sediment loads (fig. 6) shows that during flashy floods, the actual sediment load can be underestimated by 80% or more. over longer term floods, such as caused by snow-melt, calculated sediment loads using daily discharge approach those using hourly discharge (fig. 6). not knowing how sediment discharges were measured in many italian or (particularly) albanian rivers, especially during flash floods (selenica, 2004; de luque et al., 2006), the discharges shown in fig. 5 can only be considered as estimates. albanian river sediment discharge, in fact, may be significantly greater than previously reported. reliable sediment concentrations data are particularly critical to determine if river waters reached hyperpycnal concentrations. considering their generally higher sediment concentrations (fig. 4) and the probable flashiness of their floods, it seems likely that such hyperpycnal events would occur in albanian rivers rather than in italian rivers whose sediment concentrations tend to be lower (fig. 4) and flashy events are fewer. fate of sediment discharge in the adriatic extensive study of the italian continental shelf and upper continental slope indicates that, with the fluctuation of quaternary sea level, the many italian western-flowing rivers – with the possible exception of the po – have deposited a nearly continuous series of clinoform sequences (trincardi et al., 1994; correggiari et al., 2005; cattaneo et al., 2007). because of the typical cyclonic circulation within the gulf of venice, much of the discharged sediment from the alpine and apennine rivers has remained on the relatively narrow western shelf or upper slope and drifted longshore southward in response to the southflowing western adriatic current, as indicated by the distribution of hgand pb-rich sediments in the northern adriatic (compare fig. 7 with fig. 5), in part reflecting industrial pollution (dolenec et al., 1998). in the southeastern sector of the basin the picture is different, as river discharge from albania has resulted in a series of lobate deltas, the most prominent being off the semani and vjosa rivers (fouache et al., 2001). however, mg, cr, sc, ni and co distributions in the southern and central parts of the adriatic (fig. 7) also suggest that albanian river sediments have escaped to deeper waters, perhaps via turbidity currents generated during flash floods. the possible importance of such events is supported to some extent by the initial results of integrated wave-ocean-sediment numerical models. although available measurements have focused mainly on the southwestern adriatic margin, carniel et al. (2016b), employing a 1-km coupled wave-ocean-sediment model over the entire adriatic sea, found a relatively good match between water and sediment trap samples. this study, on the wake of the work by harris et al. (2008), suggested that in this transition area, given sufficient data, state-ofthe-art integrated models are capable of describing hyperpycnal sediment fluxes – both in timing and spatial distribution on the shelf and upper slope. time-averaged, vertically integrated water fluxes in the southern adriatic pit region between february 01 and may 31, 2012 (fig. 8, left) show strong transport-generated dense down-flow within the deepest region of the bari canyon. however, a relative maximum driven by a cyclonic circulation is also depicted within the sap; along the eastern coast, the prevailing direction is northward. fig. 8 (right panel) shows the vertically integrated sediment fluxes for the same time period. within the sap, the relative weight of regional cyclonic circulation, which fig. 6. comparison of daily sediment loads for small mountainous rivers in taiwan and southern california computed using daily and hourly water discharges various floods. we define flashiness of a flood as being the time between when water discharge increases from 1/3 peak discharge to peak discharge and back again to 1/3 peak discharge. in this figure, if flashiness is less than 5 hours, calculated sediment loads using daily discharge data tend to account for only ~20% of the calculated loads using hourly data. when the flashiness index approaches or exceeds 4 days (96 h), such as during snow melt, calculated sediment load based on daily discharge approaches the sediment load based on hourly discharge. inadequately monitored floods, such as flash floods that can occur in albania, may thus lead to underestimated calculated sediment discharge if it is based on daily discharge data. 112 j.d. milliman et al. controls abyssal dynamics on the lower slope and may reach hyperpycnal concentrations, leads to a semi-steady pattern lasting several weeks, and is likely capable of redistributing sediment, oxygen, nutrients and biomass throughout the entire abyssal plain (fig. 8). possible differences in water-mass character entering the sap have also been recently suggested with regard to the benthic biology. taviani et al. (2016) reported conspicuous megafaunal sessile communities, including cold-water and sponge habitats, showing an asymmetric distribution in the southern adriatic, with most diverse and abundant live corals settling the western side, especially in bari canyon (carniel et al., 2016a). these observations are hypothesized to be in response to the seasonal cascading of dense shelf water from the western shelves (carniel et al., 2016b; bonaldo et al., 2016), specifically canyons and adjacent areas, which limits excess silting and favoring the trophic web (taviani et al., 2016). this would also support the different nature and behavior of water entrainment emanating from the eastern sector, e.g., less energetic and more silty. concluding remarks although dam construction has dramatically changed the quantity of fluvial sediment entering the adriatic sea, one can still view the area as having two key point sources of incoming sediment: the northwest, fed by five alpine rivers, most notably the po, and the southeast, fed by albanian rivers, particularly the semani and vijose. the hypopycnal character of sediment input to the western adriatic, combined with line-source input of sediments derived from the italian apennine rivers, helps explain the clinoform sequences seen along the western adriatic. in fig. 7. distribution of hg and co in adriatic sea sediments (modified from: dolenec t, faganeli j, pirc s, 1998. major, minor and trace elements in surficial sediments from the open adriatic sea: a regional geochemical study. geol. croat. 51:59-73) reflect distinct sources. hg-rich sediments are located seaward of the po and other italian alpine rivers, suggesting relatively little offshore transport. in contrast, co-rich sediments, presumably derived from albanian rivers, is spread throughout much of the southern adriatic, suggesting offshore escape via hyperpycnal turbidity currents. river sediment flux and fate in the adriatic sea 113 contrast, because of periodic hyperpycnal discharge and counter-clockwise circulation, albanian river sediments are transported northward into the central parts of the adriatic sea. several areas of future research are needed if we are to more fully understand the flux and fate of terrigenous sediments in the adriatic sea. first is a better documentation of river-borne sediment. this would include analysis – or reanalysis – of existing data. it may also require obtaining new river data, especially during flash floods. knowing the difficulty to sample during a sudden flash flood, using an acoustic method might be advisable (guerrero et al., 2016). second, such a dichotomy of source and transport as is seen in the adriatic merits carefully positioned long-cores, ideally based on detailed high-resolution seismic profiles. acknowledgments we thank giuseppe morabito and luigi naselli flores, for the opportunity they gave jdm to attend the associazione italiana di oceanologia e limnologia (aiol) annual meeting. we also acknowledge the cnr short term mobility support grant that allowed jdm to work in venice. we thank katherine farnsworth for her invaluable help in preparing figs. 3 and 5, as well as t.y. lee and j.c. huang (taiwan) and j. warrick (usa) for help in gathering data used in fig. 6. we thank francesco comiti and another anonymous reviewer for their helpful comments on an earlier version of this paper and massimo guerrero for our fruitful discussions. references bethge p, 2014. hydropower struggle: dams threaten europe’s last wild rivers. spiegel online in. available from: http://www.spiegel.de/international/europe/hydropowerdams-threaten-river-wildlife-in-balkans-a-943318.html bonaldo d, benetazzo a, bergamasco a, campiani e, foglini f, sclavo m, trincardi f, carniel s, 2016. interactions among adriatic continental margin morphology, deep circulation and bedform patterns. mar. geol. 375:82-98. carniel s, benetazzo a, boero b, boldrin a, falcieri fm, langone l, sclavo m, taviani m, trincardi f, bonaldo d, 2016a. interdisciplinary investigation of offshelf transport in the southern adriatic sea: the role of bari canyon. ciesm monograph 47:145-156. carniel s, bonaldo d, benetazzo a, bergamasco a, boldrin a, falcier, fm, sclavo m, trincardi f, langone l, 2016b. offshelf fluxes across the southern adriatic margin: factors controlling dense-water-driven transport phenomena. mar. geol. 375:44-63. cattaneo a, trincardi f, asioli a, correggiari a, 2007. the western adriatic shelf clinoform: energy-limited bottomset. cont. shelf res. 27: 506-525. ciavola p, mantovani f, simeoni u, tessari u, 1999. relation between river dynamics and coastal changes in albania: an assessment integrating satellite imagery with historical data. int. j. remote sens. 20:561584. correggiari a, cattaneo a, trincardi f, 2005. the modern po delta system: lobe switiching and asymmetric prodelta growth. mar. geol. 222:49-74. de luque a, porja t, martin a, guijarro ja, alonso s, 2006. a case of severe flood over albania: a rainfall analysis from a satellite perspective. adv. geosci. 7:65-72. dolenec t, faganeli j, pirc s, 1998. major, minor and trace elfig. 8. vertically integrated water fluxes (left panel) and sediment fluxes (right panel) driven by dense water flow in the southern adriatic pit region. time-averaged between february 01 and may 31, 2012. adapted from: marine geology, vol. 375, pages 44-63, 2016, carniel s, bonaldo d, benetazzo a, bergamasco, boldrin a, falcieri fm, sclavo m, f trincardi, langone l, ‘off-shelf fluxes across the southern adriatic margin: factors controlling dense-water-driven transport phenomena’, with permission from elsevier. 114 j.d. milliman et al. ements in surficial sediments from the open adriatic sea: a regional geochemical study. geol. croat. 51:59-73. drake de, kilpack r, fischer pj, 1972. sediment transport on santa barbara-oxnard shelf, santa barbara channel, california, p. 307-332. in: j.p. swift, d.b. duane and o.h. pilkey (eds.), shelf sediment transport: processes and patterns. dowdon, hutchinson and ross, stroudsburg. eurosion, 2004. living with coastal erosion in europe; sediment and space for sustainability. part ii: maps and statistics, dg environment. fouache e, gruda g, ucaj s, nikolli p, 2001. recent geomorphological evolution of the deltas of the rivers seman and vjosa, albania. earth surf. proc. land. 26:793-802. guerrero m, latosinski f, szupiany rn, nones m, re m, gaeta mg, 2015. a sediment fluxes investigation for the 2-d modelling of large river morphodynamics. adv. water resour. 81:186-198. guerrero m, rüther n, szupiany r, haun s, baranya s, latosinski f, 2016. the acoustic properties of suspended sediment in large rivers: consequences on adcp methods applicability. water 8:13. harris ck, sherwood cr, signell rp, bever aj, warner jc, 2008. sediment dispersal in the northwestern adriatic sea. j. geophys. res. 113:c11s03. inman dl, jenkins sa, 1999. climate change and the episodicity of sediment flux of small california rivers. j. geol. 107: 251-270. kao sj, milliman jd, 2008. water and sediment discharge from small mountainous rivers, taiwan: the roles of lithology, episodic events, and human activities. j. geol. 116: 431-448. milliman jd, farnsworth kl, 2011. river discharge to the coastal ocean: a global synthesis. cambridge university press: 366 p. milliman jd, farnsworth kl, albertin cs, 1999. flux and fate of fluvial sediments leaving large islands in the east indies. j. sea res. 41:97-107. milliman jd, kao sj, 2005. hyperpycnal discharge of fluvial sediment to the ocean: impact of supertyphoon herb (1996) on taiwanese rivers. j. geol. 113:503-516. milliman jd, syvitski jpm, 1992. geomorphic/tectonic control of sediment discharge to the ocean: the importance of small mountainous rivers. j. geol. 100:525-544. mulder t, savoye b, piper d, syvitski j, 1998. the var sedimentary system: understanding holocene sediment delivery processes and their importance to the geological record. in: m. stoker, d. evans and a. cramp (eds.), geological processes on continental margins: sedimentation, mass wasting and stability. geological society of london, spec. publication 129:145-166. mulder t, syvitski jpm, 1995. turbidity currents generated at river mouths during exceptional discharges to the world oceans. j. geol. 103:285299. mulder t, syvitski jpm, migeon s, faugères jc, savoye b, 2003. marine hyperpycnal flows: initiation, behavior and related deposits. a review. mar. petrol. geol. 20:861-882. selenica a, 2004. flood potential in albania. page 94 in proc. balwois 2004 conf., ohrid, macedonia (abst.). simeoni u, pano n, ciavola p, 1997. the coastline of albania: morphology, evolution and coastal management issues. ciesm publ. 18: 561-168. syvitski jpm, kettner aj, 2007. on the flux of water and sediment into the northern adriatic sea. cont. shelf res. 27:296-308. syvitski jpm, milliman jd, 2007. geology, geography and humans battle for dominance over the delivery of fluvial sediment to the coastal ocean. j. geol. 115:1-19. taviani m, angeletti l, beuck l, campiani e, canese s, foglini f, freiwald a, montagna, p, trincardi f, 2016. on and off the beaten track: megafaunal sessile life and adriatic cascading processes. mar. geol.y 375:146-160. trincard, f, correggiari a, roveri m, 1994. late quaternary transgressive erosion and deposition in a modern epicontinental shelf: the adriatic semi-enclosed basin. geo-mar. lett. 12:41-51. waananen ao, 1969. floods of january and february 1969 in central and southern california. u.s. geological survey open file report: 223 p. warrick ja, milliman jd, 2003. hyperpycnal sediment discharge from semiarid southern california rivers: implications for coastal sediment budgets. geology 31:781-784. layout 1 introduction the european union’s water framework directive (wfd; european union 2000) requires the management of surface water body in order to achieve (or maintain) at least good chemical and ecological status. good ecological status is defined as a slight deviation from the ecological status normally associated with the same surface water body type under undisturbed conditions. according to annex 5 of the wfd, the ecological status should be evaluated using the so-called biological quality elements (bqes), which include fish, invertebrates, macrophytes and phytobenthos, and phytoplankton. for each bqe, the value of a quality index should be compared to a reference value indicating minimal human impact to obtain an ecological quality ratio. macrophytes and phytobenthos are components of freshwater flora and are included by annex 5 of the wfd in a single bqe, but they are evaluated separately, as they can respond differently to human impact, in relation to site-specific conditions (schneider et al., 2012; kelly et al., 2016). within phytobenthos, diatoms are considered the most representative component and are commonly used for the evaluation of the ecological status of both lotic and lentic waterbodies. since the pioneering work by sladecek (1984), diatoms have been largely used for the evaluation of river quality. among the methods used in europe, kelly et al. (2014) distinguish three types: indices based on the weighted average equation of zelinka and marvan (1961) and optimised against a stressor gradient, indices based on the relative proportion of taxa associated with unimpacted and impacted conditions, and multimetric indices based on a combination of these approaches. in the oldest indices, species were grouped on the basis of their tolerance or sensitivity to pollution, and river quality was estimated on the basis of the proportion of tolerant and sensitive taxa. more recent indices, such as ips (cemagref, 1982), trophic diatom index (tdi, kelly and whitton, 1995) and the diatom biological index (bdi, coste et al., 2009), are based on a weighted average equation (zelinka and marvan, 1961) and make use of ecological information for each species, obtained from a calibration data set. during the calibration, each species is assigned a value indicating the sensitivity or tolerance of that species to a specific form of pollution, and an optional value indicating the indicator value of that species. the development of specific indices for the evaluation of lake ecological quality (rott, 1999; schaumburg et al., 2004; flemish environment agency, 2009; bennion et al., 2014) is more recent. most indices using diatoms for evaluating the ecological quality of lakes are based on the trophic preferences of single species. this approach may cause two difficulties: on one hand, the large number of diatom species requires intensive training of the personnel performing the analysis and species misidentification can occur; on the other hand, a given index cannot be used when the sample is rich in species that are not included in advances in oceanography and limnology, 2018; 9(1): short note doi: 10.4081/aiol.2018.7389 this work is licensed under a creative commons attribution-noncommercial 4.0 international license (cc by-nc 4.0). using benthic diatoms for estimating lake ecological quality: comparing different taxonomic resolution aldo marchetto,* tommaso sforzi cnr-ise institute of ecosystems study, largo tonolli 50, 28922 verbania pallanza, vb, italy *corresponding author: a.marchetto@ise.cnr.it abstract the water framework directive asks to all member states of the european union to classify the ecological quality of significant waterbodies on the basis of the biological communities they host. one of the biological communities that must be used for the ecological quality assessment is the periphytic community, mainly composed by diatoms. in italy, diatom-based lake quality assessment is performed using a specific index, named epi-l, based on the method of weighted averages. for each species, a trophic score and an indicator weight were calculated. in order to reduce the complexity of the lake quality assessment, we calibrated a variant of epi-l, using diatoms genera instead of species, and we compared the performance of these two variants in terms of correlation with the nutrient level and of different classification of each lake. key words: diatoms; wfd; lakes. received: march 2018. accepted: june 2018. no nco mm er cia l u se on ly a. marchetto and t. sforzi2 the list that was defined when setting up the index. most indices were developed in the 1990s, when diatom species were grouped in large, heterogeneous genera. in the last decades, a large revision of diatoms taxonomy has occurred (round et al., 1990), splitting genera with a large number of species into smaller units, morphologically more homogenous. although it can be expected that species belonging to smaller and probably more genetically homogenous genera would also have closer ecological preferences, it is possible that the new genera still include species with large ecological differences. in this paper, we test the possibility to develop an index using diatom genera instead of species, for reducing the effort and the cost needed for the routine ecological classification of lakes and reservoirs and for increasing the number of waterbodies for which the index can be applied, including those that are rich in diatoms belonging to genera included in the list, even if their species are not the same used for the index calibration. for this purpose, we develop a quality index based on diatom genera and we compare it with an existing index based on diatom species, namely the index epi-l used in italy (marchetto et al., 2013), and we discuss their relative ability to classify lentic waterbodies in quality classes. methods during the period 2010-2017, a total of 108 diatom samples were collected from 64 lakes in italy (tab. 1), following a common protocol (buzzi et al., 2014), mainly based on cen-en standard 13946 (european committee for standardization, 2014). most of the samples were collected on permanently submerged stones using a toothbrush, while 23 samples were collected on macrophyte stems. twenty lakes were sampled more than once (tab. 1). samples were digested in h2o2 and hcl, mounted in naphrax and a minimum of 400 diatom valves were identified under optical microscope at 1000x. results from a previous european project (emerge: “european mountain lake ecosystems: regionalisation, diagnostic & socio-economic evaluation”) allowed the addition of further 16 epilithic samples, collected in 2001 in high mountain lakes with moderate to high alkalinity (higher than 0.2 meq l–1) with a very similar protocol (marchetto et al., 2009). to avoid the overrepresentation of more intensively sampled lakes, in this study we used the mean species abundance for each lake where several samples were available. the complete data set is available at: http://www.ise.cnr.it/products/datasets. the diatom species counts were entered into omnidia version 5.3 (lecointe et al., 1993), a diatom database and indices calculation tool. the following indices, commonly used for lake or river quality assessment in europe, were calculated: biological diatom index (bdi; coste, 1999), pollution sensitivity index (ips: cemagref, 1982), saprobic index (si; rott et al., 1997), trophic index (ti; rott et al., 1999), and the trophic diatom index (tdi; kelly and whitton, 1995). the epi-l index (marchetto et al., 2013) was calculated using the tool available at: http://www.ise.cnr.it/it/wfd species occurring with a higher abundance than 1% in less than 3 lakes or never reaching a minimum abundance of 3% in any sample were discarded. the epi-l index was used for testing the effect of different taxonomical resolution. for this purpose, the index was recalibrated using the present data set, and excluding species occurring with a higher abundance than 1% in less than 3 lakes or never reaching a minimum abundance of 3% in any sample. the calibration follows the procedure used by marchetto et al. (2013) and summarised here. for each ith remaining species, a trophic weight (p’) was obtained by the average of the logarithm of the epilimnetic total phosphorus concentration (tp, in µg l–1), weighted by the abundance of that species in each j-th lake (a). (eq. 1) the indicator value (v) was obtained as the inverse of the average of the squared differences between the trophic weight of the species and the epilimnetic total phosphorus concentration in each j-th lake, weighted by the abundance of that species in the j-th lake itself. indicator values higher than 30 were replaced with 30. (eq. 2) the re-calibrated epi-l index (epi-l’species) was then calculated on the basis of the relative abundance i-th species using the following formula: (eq. 3) the index value was not calculated for those lakes for which the sum of the relative abundance of the discarded species was higher than 30% the trophic value of each species was then corrected using a linear regression of the newly calculated index values for all lakes against the original epi-l values. (eq. 4) the slope and intercept values of the regression equation were used to rescale the trophic weights: (eq. 5) rescaling species optima (marchetto, 1994) was used to obtain epi-l’ values close to the original epi-l in order no nco mm er cia l u se on ly diatom taxonomic resolution 3 tab. 1. list of the considered lakes. lake longitude (°e) latitude (°n) altitude (m asl) alkalinity mean tp no. of samples albano 41.745 12.676 293 h 20 1 alserio 45.785 9.213 280 h 8 1 annone (w basin) 45.817 9.333 224 h 29 1 antholzer see 46.886 12.166 1640 m 4 1 antrona 46.054 8.091 1083 m 5 1 aplanersee 46.449 10.874 2367 m 3 1 avigliana 45.065 7.387 352 h 12 3 baratz 40.683 8.224 32 h 60 1 bidighinzu 40.557 8.662 330 h 100 1 bilancino 43.978 11.281 252 h 14 1 boden inferiore 46.442 8.453 2334 m 4 1 boden superiore 46.439 8.453 2343 m 4 1 bolsena 42.583 11.933 305 h 22 4 bracciano 42.117 12.233 164 h 16 1 caldonazzo 46.033 11.243 450 h 7 1 campo 46.129 8.131 2293 m 4 1 candia 45.324 7.912 227 h 16 3 capezzone 45.941 8.210 2100 m 4 1 cavazzo 46.333 13.077 195 h 3 2 cavedine 46.000 10.951 241 h 17 1 chiusi 43.055 11.962 251 h 32 1 cuga 40.613 8.464 642 m 24 1 endine 45.778 9.938 334 h 15 1 fusine inferiore 46.482 13.669 924 h 3 1 fusine superiore 46.477 13.668 929 h 4 1 garda 45.667 10.700 133 h 9 1 garlate 45.821 9.406 198 h 12 1 grande di monticchio 40.930 15.610 656 h 87 1 grünsee 46.609 12.009 2043 h 2 1 haidersee 46.757 10.532 1449 m 13 1 karersee 46.426 10.703 1519 h 3 1 klammsee 46.982 12.128 2258 h 4 1 kratzbergersee 46.705 11.286 2119 m 4 1 ledro 45.878 10.751 655 h 9 1 levico 46.014 11.278 440 h 5 1 liscia 40.994 9.244 178 m 29 1 lungo 42.475 12.849 371 h 48 2 maggiore 45.950 8.667 194 m 7 2 martignano 42.115 12.303 207 h 15 3 massaciuccoli 43.833 10.333 2 h 21 1 matogno 46.251 8.401 2067 m 4 1 mergozzo 45.956 8.463 194 m 4 2 mezzano 42.631 11.765 455 h 17 1 mezzola 46.199 9.441 199 m 22 2 milchsee 46.726 11.072 2540 m 3 1 molveno 46.123 10.959 823 m 4 1 montepulciano 43.090 11.920 249 h 90 1 monterosi 42.205 12.294 237 h 55 1 morasco 46.423 8.395 1815 h 3 1 nemi 41.714 12.703 318 h 27 1 orta 45.817 8.400 290 m 5 1 paione inferiore 46.169 8.191 2002 m 3 1 paione medio 46.172 8.192 2147 m 6 1 palù 46.199 9.868 1925 m 5 2 to be continued on next page no nco mm er cia l u se on ly a. marchetto and t. sforzi4 to avoid differences in lake classification between epi-l and epi-l’ and to avoid the need of defining new boundaries between the quality classes for the revised index. the same procedure was then used considering diatom genera instead of species to obtain a second index, named epi-l’genera. results all calculated diatom indices resulted significantly correlated to the trophic gradient, expressed as the mean epilimnetic total phosphorus concentration (fig. 1), but the high correlation between epi-l and tp is an artefact, due to the use in this paper of the same samples used for the calibration of the index. lakes with low phosphorous concentration can be distinguished from lakes with high concentration on the basis of the epi-l index. on the contrary, in spite of the good correlation, for all other indices there is a large overlap between the values calculated for lakes with high and low trophic status. for this reason, epi-l was selected for the quality assessment of italian lakes, and is also used in this study. trophic weights and indicator values were obtained for 90 species and are reported in tab. s1 in the supplementary material, after rescaling (m=1.0761, n=0.0472). the new index (epi-l’species) resulted strongly correlated (r=0.874, p<0.001) with the trophic gradient, expressed as the decimal logarithm of the mean concentration of total phosphorus (fig. 2a). repeating the calibration for 34 genera, the index (epi-l’genera) also resulted strongly correlated (r=0.683, p<0.001) with the trophic gradient, expressed as the decimal logarithm of the mean concentration of total phosphorus, but some outliers were evident (fig. 2b). indeed, for most genera, the trophic scores of the species lye in a relatively small range (fig. 3), so that the use of an index based on lower taxonomic resolution should be possible. however, this was not the case for the genera achnanthidium, discostella, encyonema, eunotia, fragilaria and pantocsekiella (fig. 3). we suspected that the presence in these genera of species with outlying trophic score could lead to a weaker performance of the table 1. continued from previous page. lake longitude (°e) latitude (°n) altitude (m asl) alkalinity mean tp no. of samples panelatte 46.203 8.458 2063 m 7 1 paterno 42.382 13.014 617 h 40 1 pattada 40.575 9.167 561 m 50 1 piano 46.037 9.162 276 h 14 1 piccolo di avigliana 45.054 7.392 356 h 70 1 piccolo di monticchio 40.932 15.619 658 h 23 1 piediluco 42.529 12.751 368 h 33 2 pojala 46.329 8.335 2305 m 5 1 posada 40.639 9.608 43 m 45 1 predil 46.417 13.567 965 h 3 1 pusiano 45.802 9.273 259 h 11 2 ragogna 46.175 13.003 188 h 13 2 ripasottile 42.475 12.815 371 h 60 2 salto 42.279 13.024 535 h 49 2 scanno 41.923 13.864 922 h 21 2 segrino 45.829 9.267 374 h 11 1 sirio 45.485 7.885 271 h 18 3 sos canales 40.555 9.313 711 m 28 1 südlichter kofferrastersee 46.576 10.940 2405 m 6 1 tenno 45.939 12.452 570 h 3 1 timmelsschwarzsee 46.928 11.163 2514 m 3 1 toblino 46.063 10.967 245 h 24 1 trasimeno 43.150 12.100 259 h 60 2 turano 42.232 12.941 540 h 62 3 vico 42.317 12.167 507 h 21 1 viverone 45.401 8.051 230 h 30 3 h, alkalinity >1 mmolc l–1; m, alkalinity between 0.2 and 1 mmolc l–1; tp, total phosphorus (µg l–1). no nco mm er cia l u se on ly diatom taxonomic resolution 5 index when the taxonomic resolution would be reduced. for this reason, beside an index based on diatom species (epi-lspecies) and one based on diatom genera (epi-lgenera), we developed a hybrid index based on more homogeneous group (epi-ltaxa). for calibrating this latter index, we separated the relative abundance of achanthidium lineare w.smith, discostella stelligera (cleve & grunow) houk & klee, encyonema caespitosum kützing and e. ventricosum (c.agardh) grunow, eunotia exigua (brébisson ex kützing) rabenhorst and pantocsekiella ocellata (pantocsek) k.t.kiss & e.ács from the rest of their respective genera. in the case of fragilaria, we considered three groups: the first formed by f. crotonensis kitton, the second by f. tenera (w.smith) lange-bertalot and f. tenera var. nanana (lange-bertalot) langebertalot & s.ulrich, and the third including all other fragilaria species. trophic weights and indicator values were then obtained for 40 taxa (genera and species groups) and are reported in table s2, after rescaling (m=1.2108, n=0.1981). the correlation of this index, named epi-l’taxa, fig. 1. relationship between some selected diatom-based quality index and the average annual total phosphorus concentration for the lakes used in this study. no nco mm er cia l u se on ly a. marchetto and t. sforzi6 with the trophic gradient was only slightly stronger (r=0.689, p<0.001) than the one based on genera (fig. 2c). comparing epi-l’taxa with the original epi-l, their values compare well, with a median absolute difference of 0.02 (fig. 2d). discussion in our database, the use of metrics developed for rivers for the assessment of ecological status in lakes resulted in slight correlation with the trophic gradient, probably because some diatom species do have distinct preferences for lakes over rivers (kelly et al., 2014). among the metrics specifically developed for lakes, l-tdi (bennion et al., 2014) could not be correctly applied to our data set, as in no lakes the abundances of the species included in the l-tdi species list accounted for more than 70% of the total diatom counts. therefore, the correlation between ltdi and the logarithm of the total phosphorus concentration was low (r=0.44). on the contrary, rott’s (1999) ti index was well correlated with the trophic gradient (r=0.64), but it could only be used to evaluate the ecological quality of 52 out of 80 lakes in our database. for this reason, marchetto et al. (2013) developed a specific index for italian lakes and reservoirs (epi-l), in order to have a species list reflecting the composition of diatom assemblages in italy and a good relationship with the trophic gradient in these lakes, which can be applied to 73 lakes in our data set. however, we expect that in a large-scale monitoring of all italian lakes and reservoirs, some species not included in the epi-l list will be found, making impossible to assess the ecological quality of some lakes or reservoirs. to allow the classification of those water bodies, we propose to use a revised index that does not require diatom determination at the species level. the indices based on species and on taxa was able to classify 74 out of 80 lakes, while the index based on genera was able to classify 79 out of 80 lakes. our results substantially confirm the finding of bennet et al. (2014), who discussed the performance of diatom indices using different taxonomic resolution. they found large performance differences between speciesand fig. 2. scatter plot of mean epi-l’species (a), epi-l’genera (b) and epi-l’taxa (c) vs the average annual total phosphorus concentration, and comparison of epi-l’taxa with the original epi-l for the lakes used in this study (d). no nco mm er cia l u se on ly diatom taxonomic resolution 7 genus-based indices calibrated along an acidity gradient, however, the differences were small when comparing indices calibrated along a trophic gradient. following the wfd, the boundaries between quality classes should be defined through intercalibration exercises, to assure that in the same ecoregion class boundaries are shared by all member states. in the case of lake diatoms, the intercalibration exercise (kelly et al., 2014) was performed comparing the methods used in different countries with a common metric, namely the trophic index (rott, 1999), which requires diatom counts at the species level. for this reason, if the epi-lgenera or epi-ltaxa will be used for wfd-compliant ecological assessment, a species-level diatom determination is still needed for the data set used for the intercalibration exercise. based on the intercalibration exercise (marchetto, 2014), the boundaries between the “high” and “good” quality classes were set to an epi-l value of 1.702 for deep lakes and 1.845 for shallow lakes, while the boundaries between “good” and “moderate” ecological quality was set to 1.135 for deep lakes and 1.230 for shallow lakes. in this study, most lakes fell in the same quality class using either epi-l’species, epi-l’genera or epi-l’ taxa. differences in lake classification between epi-l’species, and epi-l’ taxa and between epi-l’genera and epi-l’ taxa were only found for two lakes, for which the index value lied close to the boundary value. however, the number of classification mismatches was higher (6) in the case of epi-l’genera. these results are similar to those obtained by bigler et al. (2010) in a similar study: they also found that reducing taxonomic resolution within the achnanthidium minutissimum species complex from a series of subgroups to a single group led to changes in lake quality classification using the ips index for a small number of lakes, only in cases when the index value was close to the class boundary. conclusions diatom-based indices are widely used in europe for assessing lake ecological quality. most of them are based on a list of indicator values assigned to diatom species. they cannot be used when a significant proportion of diatom found in a given lake are not included in their species list. to reduce this problem, we tested the possibility to use an index based on diatom genera. however, we found that the reduction in taxonomic resolution led to an increase in classification mismatches. the results of this exercise indicate that the index based on genera can be improved in classification ability splitting some genera, when the species belonging to them have markedly different trophic preferences. the resulting index seems to compare with an index requiring diatom determination to the species level. finally, if adequate rescaling of the trophic scores is applied, we propose that the genera trophic scores and indicator values can be used together with the species ones, when individual specimen cannot be assigned to a given species, or when their species is not included in the species list of the original index. acknowledgments we warmly acknowledge dr. mattia azzella for providing samples of the volcanic lakes in central italy and the following data providers: renate abler, berta thaler (provincial environmental agency of bolzano), arianna macor, raffaella zorza (regional environmental agency of friuli venezia giulia), simona musazzi (cnrfig. 3. distribution of the trophic scores (p’) for species belonging to the same genus. no nco mm er cia l u se on ly a. marchetto and t. sforzi8 ise), stefania balsamo (italian institute for environmental protection and research), laura mancini, stefania marcheggiani, camilla puccinelli, claudia vendetti (national institute of health), chiara agostinelli, andrea beghi, fabio buzzi, elisa carena, marco fioravanti, riccardo formenti, matteo galbiati, filippo galimberti (regional environmental agency of lombardia), pierluigi fogliati, arianna nicola (regional environmental agency of piemonte), silvia costaraoss, sabrina pozzi, paola testa (provincial environmental agency of trento), susanna cavalieri, federica cimoli, giovanna marchi, chiara rustighi (regional environmental agency of toscana), stefania bracchi, isa crescentini, valentina della bella, margerita di brizio, rosalba padula, egiziana rinaldi (regional environmental agency of umbria). high mountain diatom samples were collected within the emerge project (“european mountain lake ecosystems: regionalisation, diagnostic & socioeconomic evaluation”), funded by the european union (contract evk1-ct-1999-00032) under the 5th framework programme. references bennett jr, sisson dr, smol jp, cumming bf, possingham hp, buckley ym, 2014. optimizing taxonomic resolution and sampling effort to design cost-effective ecological models for environmental assessment. j. appl. ecol. 51:1722-1732. bennion h, kelly mg, juggins s, yallop ml, burgess a, jamieson bj, krokowski j, 2014. assessment of ecological status in uk lakes using benthic diatoms. freshwat. sci. 33:6 39-654. bigler c, gälman v, renberg i, 2010. numerical simulations suggest that counting sums and taxonomic resolution of diatom analyses to determine ips pollution and acid acidity indices can be reduced. j appl phycol 22:541-548. buzzi f, mancini l, vendetti c, puccinelli c, marcheggiani s, marchetto a, 2014. [protocollo di campionamento ed analisi delle diatomee bentoniche dei laghi e degli invasi], p. 215233. in: s. balzamo and c. martone (eds.), [metodi biologici per le acque superficiali interne].[book in italian]. ispra, manuali e linee guida 111/2014. cemagref, 1982. [etude des méthodes biologiques d’appréciation quantitative de la qualité des eaux, in bassin rhône-mediterrannée-corse].[book in french]. csiro: pp. 218. european committee for standardization, 2014. water quality guidance for the routine sampling and preparation of benthic diatoms from rivers and lakes. european standard 13946. coste m, boutry s, tison-rosebery j, delmas f. 2009. improvements of the biological diatom index (bdi): description and efficiency of the new version (bdi-2006). ecol. indic. 9:621-650. descy jp. 1979. a new approach to water quality estimation using diatoms. nova hedwig 64:305-323. european union, 2000. directive 2000/60/ec of the european parliament and of the council of 23 october 2000 establishing a framework for community action in the field of water policy. off. j. (l 327), 22 (2000), p. 12 flemish environment agency, 2009. biological assessment of the natural, heavily modified and artificial surface water bodies in flanders according to the european water framework directive. vlaamse milieumaatschappij, erembodegem, brussels. kelly m g, whitton ba, 1995. the trophic diatom index: a new index for monitoring eutrophication in rivers. j. appl. phycol. 7:433-444. kelly m, urbanic g, acs e, bennion h, bertrin v, burgess a, denys l, gottschalk s, kahlert m, karjalainen sm, kennedy b, kosi g, marchetto a, morin s, picinskafałtynowicz j, poikane s, rosebery j, schoenfelder i, schoenfelder j, varbiro g. 2014. comparing aspirations: intercalibration of ecological status concepts across european lakes for littoral diatoms. hydrobiologia 734:125141. kelly mg, birk s, willby nj, denys l, drakare s, kahlert m, karjalainen sm, marchetto a, pitt j-a, urbanič g, poikane s, 2016. redundancy in the ecological assessment of lakes: are phytoplankton, macrophytes and phytobenthos all necessary? sci. total environ. 568:594-602. lecointe c, coste m, prygiel j, 1993. ‘omnidia’: software for taxonomy, calculation of diatom indices and inventories management. hydrobiologia 269/270:509-513. lenoir a, coste m, 1996. development of a practical diatom index of overall water quality applicable to the french national water board network. in: b.a. whitton and e. rott (eds.), use of algae for monitoring rivers ii. institute of botanic, university of innsbruck. marchetto a, 1994. rescaling species optima estimated by weighted averaging. j. paleolimnol. 12:155-162. marchetto a, 2014. report on fitting the italian national method for the evaluation of the ecological quality of lake waterbodies using benthic diatoms (epi-l) in the “phytobenthos cross-gig” intercalibration exercise. report cnr ise, 01.14, 18 pp. http://www.ise.cnr.it/it/wfd marchetto a, rogora m, boggero a, musazzi s, lami a, lotter af, tolotti m, thies h, psenner r, massaferro j, barbieri a, 2009. response of alpine lakes to major environmental gradients, as detected through planktonic, benthic and sedimentary assemblages. advanc. limnol. 62:419-440. marchetto a, agostinelli c, alber r, beghi a, balsamo s, bracchi s, buzzi f, carena e, cavalieri s, cimoli f, costaraoss s, crescentini i, della bella v, di brizio m, fioravanti m, i fogliati p, formenti r, galbiati m, galimberti f, macor a, mancini l, marcheggiani s, marchi g, musazzi s, nicola a, padula r, pozzi s, puccinelli c, rinaldi e, rustighi c, testa p, thaler b, vendetti c, zorza r, 2013. [indice per valutazione della qualità delle acque lacustri italiane a partire dalle diatomee epifitiche ed epilitiche (epi-l)], p. 75-92. in: a. marchetto (ed.), [indici per la valutazione della qualità ecologica dei laghi].[report in italian]. report cnr-ise 02-13. available from: http://www.ise.cnr.it/it/wfd round fe, crawford rm, mann dg, 1990. diatoms: biology and morphology of the genera. cambridge university press, p. 747. no nco mm er cia l u se on ly diatom taxonomic resolution 9 rott e, hofmann g, pall k, pfister p, pipp e, 1997. indikationslisten für aufwuchsalgen. teil 1. saprobielleindikation. bundesministeriumfür landund forstwirtschaft, wien. 1-73. rott e, 1999. [teil 2: trophieindikation sowie geochemische präferenz; taxonomische und toxikologische anmerkungen].[book in german]. indikationslisten für aufwuchsalgen in österreichischen fliessgewässern, wien. schaumburg j, schranz c, hofmann g, stelzer d, schneider s, schmedtje u, 2004. macrophytes and phytobenthos as indicators of ecological status in german lakes: a contribution to the implementation of the water framework directive. limnologica 34:302-314. schneider sc, lawniczak ae, picińska-faltynowicz j, szoszkiewicz k, 2012, do macrophytes, diatoms and nondiatom benthic algae give redundant information? results from a case study in poland. limnologica 42: 04-211. sladecek v, 1986. diatoms as indicators of organic pollution. acta hydrochim. hydrobiol. 14:555-566. zelinka m, marvan p, 1961. [zur präzisierung der biologischen klassifikation der reinheit flieβender gewässer].[article in german]. arch. hydrobiol. 57:389-407. no nco mm er cia l u se on ly layout 1 introduction research of algae in serbia began over 130 years ago, with the publication of ‘fragmenta phycologiae bosniaco-serbicae’ by scharschmidt (1883), who listed 46 algal species. the species were determined in samples of washed silt from herbarium specimens of the waterwheel plant (aldrovanda vesiculosa (l.)) collected by the famous lecturer and botanist, doctor josif pančić (blaženčić, 1986). over 50 species, mostly belonging to the green algae and cyanobacteria, were identified by magnus, simić and katić during the late 19th and early 20th centuries (milovanović, 1949). another notable name in this initial period of algal research was nedeljko košanin (blaženčić, 1986). world war i prevented further investigations of algae and cyanobacteria in serbia, however, during the 1930s and 1940s exploration continued in the form of complex hydrobiological studies. a bibliographic review of the few serbian algological studies until 1947 was presented by milovanović (1949). this period is also important because it can be seen as the cradle of almost all of the subsequent trends in algal research (blaženčić, 1986). after world war ii, favorable conditions for more advanced, and continuous studies in this scientific area were created. during this period, the spatial distribution, seasonal dynamics and ecology of algae and cyanobacteria were investigated in springs, streams, rivers, ponds, swamps, canals, lakes and mineral waters, i.e., in most aquatic ecosystems in serbia, which resulted in a wealth of data (blaženčić, 1986). since the 1970s the field of research has been expanding as methods and techniques have been modernized. an increasing number of hydrobiologists, microbiologists and botanists are providing their contributions to the study of cyanobacteria in serbia (blaženčić, 1985). undoubtedly the greatest contribution to the knowledge of algae and cyanobacteria has been given through the meticulous advances in oceanography and limnology, 2017; 8(1): 153-160 article doi: 10.4081/aiol.2017.6360 this work is licensed under a creative commons attribution-noncommercial 4.0 international license (cc by-nc 4.0). review of 130 years of research on cyanobacteria in aquatic ecosystems in serbia presented in a serbian cyanobacterial database zorica svirčev,1,2 nada tokodi,1* damjana drobac1 1department of biology and ecology, faculty of sciences, university of novi sad, trg dositeja obradovića 3, 21000 novi sad, serbia; 2department of biochemistry and pharmacy, faculty of science and engineering, åbo akademi university, tykistökatu 6 a, 20520 turku, finland *corresponding author: nada.tokodi@dbe.uns.ac.rs abstract the presence of toxic cyanobacteria in aquatic ecosystems in the territory of the republic of serbia was surveyed over a period of several decades. increasing attention is being paid to some negative consequences that may be caused by these microorganisms. information from available literary sources regarding the distribution and frequency of cyanobacteria and their toxins over a period of 130 years, together with the effects on humans and wildlife in aquatic ecosystems, were gathered and incorporated into a serbian cyanobacterial database created for the cyanocost action. this database encompasses information on 65 aquatic ecosystems, including rivers, lakes, ponds, canals, irrigation reservoirs, reservoirs used for drinking water supply and reservoirs used for other purposes. cyanobacterial blooms were found in almost 80% of the investigated aquatic ecosystems. the analysis of the research showed the presence of more than 70 species, including blooms of 24 species from 13 genera. five species of cyanobacteria: microcystis aeruginosa, aphanizomenon flos-aquae, planktothrix agardhii, microcystis flos-aquae and planktothrix rubescens frequently formed blooms in the investigated waterbodies and cyanotoxins were also detected in some of them, which had certain negative effects. here, we present an overview of data contained in the serbian cyanobacterial database, concerning cyanobacterial distribution, cyanotoxin production and associated biological effects in different types of water bodies from the republic of serbia. also, recent important and major cases of cyanobacterial blooming in reservoirs used for drinking water supply at vrutci and ćelije, the aleksandrovac irrigation reservoir, the ponjavica river and lake palić, including systematic research on the lake ludoš and few fishponds are further described. it can be concluded that cyanobacteria and cyanotoxins are omnipresent in different water bodies throughout the republic of serbia. for these reasons it is imperative to continue with the monitoring of cyanobacteria and cyanotoxins, as well as to continuously supplement the established database with new information. the serbian cyanobacterial database represents a treasury of information on cyanobacteria and their toxins, and serves as a model for other countries in the region and beyond. key words: cyanobacterial blooms; cyanotoxins; database; historical overview. received: 25 october 2016. accepted: 23 january 2017. no nco mm er cia l u se on ly z. svirčev et al.154 work of prof. jelena blaženčić who performed systematic analysis and assessment of numerous aquatic ecosystems in serbia. most recently, available data on the presence of cyanobacteria and their toxins in aquatic ecosystems in the republic of serbia were gathered and incorporated into the serbian cyanobacterial database (scdb) (https://cloud.pmf.uns.ac.rs/index.php/s/v6tervcvcauaxqn), as a result of the cyanocost action. the objective of the present paper was to analyze, present and supplement available data from scdb. data overview data on cyanobacterial occurrence, cyanotoxin production and the possible impact on humans and aquatic ecosystems in the republic of serbia were evaluated in over 70 data sources (tab. 1). a comprehensive review of research articles, project reports, conference abstracts, dissertations, books, annual and news reports was performed, and collected data were entered into the scdb. the analysis of the research showed that 65 different aquatic ecosystems had been investigated during the last 130 years. the microalgal and cyanobacterial research summarized in the scdb consisted of over 250 analyses, of which most (about 150) were conducted since the year 2000. these analyses included determination of cyanobacteria, cyanotoxin detection and the documentation of associated biological effects and health incidents (svirčev et al., 2014a). qualitative determination of cyanobacteria showed the presence of more than 70 species in the investigated aquatic ecosystems, with 24 species from 13 genera being recorded in blooms, the most frequent being microcystis aeruginosa (kützing), aphanizomenon flos-aquae ralfs ex bornet & flahaul, planktothrix agardhii (gomont) k. anagnostidis & j. komárek, microcystis flos-aquae (wittrock) kirchner and planktothrix rubescens (de candolle ex gomont) k. anagnostidis & j. komárek (svirčev et al., 2014a). all the latin genera and species names are taken from the original publications. for the currently accepted nomenclature of the species cited in this paper, refer to the supplementary tab. 1. cyanobacterial blooms were found in almost 80% of the investigated aquatic ecosystems. this paper summarizes their distribution, cyanotoxin production and associated biological effects in different types of waterbodies from the republic of serbia. canals the canals in the vojvodina region (kralja aleksandra canal and kralja petra canal) were explored during the 1930s, when a bloom of m. aeruginosa was recorded (protić, 1935). the cyanobacterium anabaena flos-aquae (dolichospermum flos-aquae (brébisson ex bornet & flahault) p. wacklin, l. hoffmann & j. komárek) also formed a mass occurrence in the kralja petra canal (protić, 1936). recently, most research was done on the canal danube-tisa-danube system where the presence of a common cyanotoxin, microcystin (mc), was detected. the highest concentrations (up to 347 µg l–1 mc-lr equivalents) were found in the autumn of 2006 at the locality of bačko gradište where the cyanobacterial species a. flos-aquae, aph. flos-aquae, m. aeruginosa, m. flosaquae and oscillatoria agardhii (planktothrix agardhii) occurred (simeunović, 2009). ponds very limited data are available on the occurrence of cyanobacteria and cyanotoxins in marsh-wetland ecosystems (jakovljević and stanković, 1931-1932; milovanović, 1970; pujin et al., 1987; maslać et al., 1992; subakov-simić et al., 2004; fužinato et al., 2010; cvijan and fužinato, 2011, 2012). recent data from 2006 showed the first ocurrence of the invasive and potentially toxic cyanobacterial species cylindrospermopsis raciborskii (woloszynska) seenayya & subba raju in slatina, a salt marsh pond (cvijan and fužinato, 2012). the presence of the cyanobacterial species m. aeruginosa in rakina bara (jakovljević and stanković, 1931-1932), aph. flos-aquae in carska bara (pujin et al., 1987) and arthrospira tab. 1. overview of the data in serbian cyanobacterial database (1930-2012). location cyanobacteria cyanotoxins biological effects reference canals (4) over 70 species found, microcystin (mc) artemia salina bioassay over 70 literature sources: ponds (7) frequently blooming: analyses: fish histopathology peer-reviewed papers (24) rivers (13) microcystis aeruginosa ppi animal mortality international documents fishponds (8) aphanizomenon flos-aquae elisa epidemiological survey (english) (10) reservoirs for irrigation (11) planktothrix agardhii hplc national documents lakes (6) microcystis flos-aquae detection in: (non-english) (40) reservoirs for drinking water planktothrix rubescens water newspaper/internet reports (3) supply (12) soil own document (1) reservoirs with other purposes (4) -plant tissues fish tissues no nco mm er cia l u se on ly research on cyanobacteria in serbia 155 fusiformis (voronikhin) j. komárek & j.w.g. lund in slatina (fužinato et al., 2010) causes some concerns as these species are known cyanotoxin-producers. rivers interestingly, there are a large number of data on cyanobacterial blooming in river ecosystems. one of the most publicized incidents happened in 2009 when the mortalities of fish and hundreds of cows and pigs which drank water from ponjavica, the river near the town of pančevo, occurred. the detected cyanobacterial species belonged to the genera anabaena, aphanizomenon, aphanocapsa, aphanothece, chroococcus, cylindrospermopsis, geitlerinema, jaaginema, limnothrix, microcystis, phormidium, planktothrix and raphidiopsis. during 2008 and 2009 in the ponjavica river, mcs (concentrations given as mc-lr equivalents in elisa) were detected in the water (up to 4.84 mg l–1), sediment (5.7 mg per 100 g), macrophytes (up to 5.0 mg per 100 g) and fish (up to 3.3 mg per 100 g), and the dominance of the invasive cyanobacterium cyl. raciborskii was found (karadžić, 2011; natić, 2012; karadžić et al., 2013). mcs may have contributed to the deaths of animals, however, the exact cause of death has not been determined. mcs were also found in rivers krivaja, tamiš, tisa and begej, with the maximum concentrations of 80, 33, 32 and 22 µg mc-lr equivalents l–1 respectively, where aph. flos-aquae, m. flos-aquae and o. agardhii bloomed (simeunović, 2009). these and other bloom-forming cyanobacteria, such as a. flos-aquae, m. aeruginosa and oscillatoria rubescens (planktothrix rubescens), were recorded in many rivers throughout the republic of serbia (obušković, 1982, 1987, 1989, 1991; sedmak and svirčev, 2011). fishponds research on the occurrence of cyanobacteria in fishponds began with the investigations of the ečka, kolut and živača fishponds where m. aeruginosa formed blooms (milovanović and živković, 1953, 1959; milovanović, 1963), while anabaena was found blooming in the futog i fishpond (ristić et al., 1979). in kapetanski rit the presence of the invasive cyanobacterial species cyl. raciborskii was noted (ćirić et al., 2010). in the last five years, more attention has been turned to possible effects of cyanobacteria and their toxins on the quality of fish meat and cyanotoxin accumulation in fish tissues, that could consequently endanger the health of consumers. in fishponds with the code bo during 2010 and 2011, mass occurrences of cyanobacterial species aph. flos-aquae, m. aeruginosa, phormidium foveolarum (leptolyngbya foveolara (gomont) anagnostidis & komárek ), jaaginema subtilissimum (kützing ex forti) anagnostidis & komárek, pseudanabaena limnetica (lemmermann) komárek and geitlerinema amphibium (c. agardh ex gomont) anagnostidis were recorded. toxicity in an artemia salina (l.) bioassay was detected in two of the six fishponds, when the maximum concentration of mcs in water amounted to 52 mc-lr equivalents l–1 in protein phosphatase inhibition (ppi) assay and 18 µg l–1 in an enzyme-linked immunosorbent assay (elisa). besides water, mcs were found in the muscle and liver of fish, as well as in aquatic plants and sludge (world bank report, dm 4307 2011). during the summer of 2011 in another fishpond, encoded mu (fig. 1a), mass occurrences of g. amphibium, j. subtilissimum, m. aeruginosa, o. agardhii and phor. foveolarum were recorded (world bank report, dm 4307 2011). the toxicity of water samples from the fishpond was confirmed by a. salina bioassay, and the presence of mcs and saxitoxin(s) in water was identified by a ppi assay and elisa, respectively (tokodi et al., 2013, 2014; drobac, 2015; drobac et al., 2016). the highest concentrations of mcs (181 µg mc-lr equivalents l–1 in the ppi assay) were detected in a water sample from september 2011 (tokodi et al., 2014). the variant mc-rr was also detected in the muscle of fish cyprinus carpio (l.) grown in the fishponds where high cyanobacterial occurrence was detected (drobac, 2015; drobac et al., 2016). additionally, pathological alterations in the fish tissues of liver, kidneys, gills, intestine and muscle were also observed (drobac, 2015; drobac et al., 2016). the observed adverse effects and accumulation of cyanotoxins in fish tissues show that cyanobacteria and their toxins in fishponds could be hazardous to fish quality, the economy, the consumers’ health and the environment in general. reservoirs used for irrigation cyanobacteria and mcs were recorded in reservoirs in serbia which are used for irrigation. of 13 blooming species, the most frequently observed were aph. flosaquae (borkovac, bukulja, manđelos, mrtva tisa, pavlovci, provala, zobnatica), o. agardhii (borkovac, mrtva tisa, pavlovci, zobnatica), a. flos-aquae (bukulja, jegrička, pavlovci) and m. flos-aquae (koviljski rit, pavlovci) (đukić et al., 1991a, 1991b; simeunović, 2009; karadžić et al., 2010; sedmak and svirčev, 2011; svirčev et al., 2013a). the highest concentration of 280 µg l–1 mc-lr equivalents in water was recorded in 2007 in mrtva tisa (simeunović, 2009) when aph. flos-aquae and o. agardhii were abundant (fig. 1b). cyl. raciborskii was detected in september 2010 in the aleksandrovac reservoir which is used for irrigation (simić et al., 2011). an extensive fish mortality in aleksandrovac occurred on 20 december 2012 and was associated with the presence of cyl. raciborskii blooming a few weeks before the incident. almost the entire fish popno nco mm er cia l u se on ly z. svirčev et al.156 ulation was killed (over 1.7 tonnes) including the species c. carpio, silurus glanis (l.), ctenopharyngodon idella (valenciennes), hypophthalmichthys molitrix (valenciennes), abramis brama (l.), carassius gibelio (bloch), aspius aspius (l.), and squalius cephalus (l.). a. salina bioassay showed high toxicity of water samples from aleksandrovac. however, the most common cyanotoxins (mcs, cylindrospermopsin, and saxitoxin) were not detected. it is possible that some other unknown or undetected toxic metabolites of this cyanobacterium were present and were a potential cause of the fish mortality in aleksandrovac (drobac, 2015; svirčev et al., 2016a). information about cyanotoxins in reservoirs used for irrigation is important because it is known that irrigation from water sources containing cyanobacteria and cyanotoxins may affect agricultural plants and lead to accumulation of these toxins. therefore, the health risks to people and animals due to the consumption of agricultural products irrigated with cyanotoxin-containing water must be taken seriously (codd et al., 1999; crush et al., 2008; saqrane et al., 2009; chen et al., 2010; drobac, 2015). lakes in gazivode lake, sjeničko lake and veliki zaton lake anabaena circinalis (dolichospermum circinale (rabenhorst ex bornet & flahault) p. wacklin, l. hoffmann & j. komárek), a. flos-aquae and o. rubescens were observed in mass occurrences (shllaku and landner, 1992; miljković et al., 2004; sedmak and svirčev, 2011). detection of mcs was performed only in the lakes palić and ludoš. research from 2005 to 2007 shows the presence of mcs in lake palić, and the highest mc concentration, 389 µg mc-lr equivalents l–1, was detected in the autumn of 2006. the most frequently blooming species was m. aeruginosa, followed by anabaena spiroides (dolichospermum spiroides (klebhan) p. wacklin, l. hoffmann & j. komárek), a. circinalis, m. flos-aquae and microcystis wesenbergii (komárek) komárek ex komárek (simeunović, 2009). in lake palić, fish mortality was observed by the author seleši (1982), and this recurred in later years as well. in 2009 there was an extensive mortality of fish with loss of over 12 tonnes of fish stocks. reasons cited included the lack of dissolved oxygen due to an excessive production of algae followed by their decay (http://www.zjzs.org.rs/page.php?id=286). furthermore, in 2012 the invasive cyl. raciborskii was also found in this lake (institute of public health of serbia, 2013). in lake ludoš, mc concentrations reached up to 604 fig. 1. cyanobacterial blooming in fishpond (code mu) (a); reservoir used for irrigation (mrtva tisa) (b); lake (lake ludoš) (c); and reservoir used for drinking water supply (vrutci) (d). no nco mm er cia l u se on ly research on cyanobacteria in serbia 157 µg mc-lr equivalents l–1 in the summer of 2006, and the species that bloomed during the study period were aph. flos-aquae, m. aeruginosa, m. flos-aquae, m. wesenbergii and p. agardhii (simeunović, 2009). research on lake ludoš during 2011 indicated not only mcs in the water, but also their accumulation in macrophytes (phragmites communis trin., typha latifolia (l.) and nymphaea elegans hook.) and in the tissues (intestine, muscles, kidney, gills and gonads) of prussian carp (c. gibelio). histopathological changes in different organs of prussian carp (liver, kidney, gills and intestines) from lake ludoš were associated with the high abundacies of potentially toxic cyanobacterial species limnothrix redekei (van goor) meffert and p. limnetica that were found in the center of the lake (fig. 1c). given that lake ludoš is a ramsar site, the stability of this aquatic ecosystem is of great and global importance (tokodi, 2016). reservoirs used for drinking water supply unlike the vojvodina region, where groundwater is used for water supply, central serbia has a large number of surface reservoirs used for drinking water supply. there are more than 20 reservoirs used as sources of drinking water, and constant mass occurrences of cyanobacteria have been observed in nine of them (svirčev et al., 2007). in the following serbian drinking water supply reservoirs: bovan, bresnica, garaši, grlište, grošnica, gruža, krajkovac and pridvorica, the emergence of the bloomforming cyanobacteria anabaena solitaria (dolichospermum solitarium (klebahn) p. wacklin, l. hoffmann & j. komárek), aph. flos-aquae, gomphosphaeria lacustris (snowella lacustris (chodat) komárek & hindák), gomphosphaeria aponina kützing, p. limnetica and m. aeruginosa has been documented, as has the presence of cyanotoxins in some of the reservoirs (sedmak and svirčev, 2011; svirčev et al., 2014a). in ćelije, the reservoir used for drinking water supply for the city of kruševac, a bloom of a. circinalis, aph. flos-aquae and m. aeruginosa was observed in 2004. mc was found in water samples from the reservoir (650 µg mc-lr l–1) and in the tap water (2.5 µg l–1) (svirčev et al., 2009). in addition to the mentioned species, aphanizomenon issatschenkoi (cuspidothrix issatschenkoi (usachev) p. rajaniemi, komárek, r. willame, p. hrouzek, k. kastovská, l. hoffmann & k. sivonen) (2001) and j. subtilissimum were detected in ćelije (2007) (svirčev et al., 2009; sedmak and svirčev, 2011). in gruža, along with a bloom of aph. flos-aquae, ultrastructural, apoptotic and necrotic changes in the liver of perch (perca fluviatilis (l.)) were observed, as well as an impact on antioxidant biomarkers (perendija et al., 2011). recently, cyanobacterial blooms of p. rubescens have occurred in vrutci reservoir used for the water supply of the city of užice (fig. 1d), where 70.000 inhabitants were potentially exposed to cyanotoxins in december 2013. based on the number of cells per ml and concentration of mcs, according to world health organization (who, 1999), water from reservoir vrutci could be classified as a high-risk water for recreation and drinking water abstraction purposes. the results from a. salina bioassay showed significant toxicity of the cyanobacterial biomass. modest fish mortality was observed during the cyanobacterial bloom, and mcs were detected in fish, including the muscle of frozen fish from 2013 which could indicate the presence of cyanobacteria even before the confirmed bloom. furthermore, a questionnaire and epidemiological results showed that health problems possibly related to cyanotoxins (diseases of digestive system, skin and subcutaneous tissue) occurred already at least two years prior to the incident. this might be a sign that the population of the city of užice could have been exposed to the cyanobacteria and cyanotoxins even two years before the observed bloom in 2013 (svirčev et al., 2016b). chronic exposure to cyanotoxins (e.g. mcs) from drinking water could present a risk factor for primary liver cancer and possibly even other types of cancer (svirčev et al., 2010; drobac, 2015). epidemiological studies conducted in serbia have revealed a significant correlation between an increased incidence of several cancers (brain; heart, mediastinum and pleural; ovarian; testicular; gastric; colorectal; retroperitoneal and peritoneal; leukemia; malignant skin melanoma; and primary liver cancer) and cyanobacterial blooms in reservoirs used for drinking water supply (svirčev et al., 2009, 2013b, 2014b; drobac, et al. 2011; drobac, 2015). reservoirs used for other purposes reservoirs used for hydropower generation have been poorly investigated. three cyanobacterial species were noted: m. aeruginosa, aph. flos-aquae and o. rubescens. only o. rubescens formed mass occurrences in investigated reservoirs (milovanović, 1973; obušković, 1983; sedmak and svirčev, 2011). conclusions based on the reviewed data from our scdb it can be concluded that cyanobacteria and cyanotoxins are omnipresent in different waterbodies throughout the republic of serbia. a systematic review and meta-analyses of the available literature is useful for an understanding of cyanobacterial biodiversity in serbian waters. some information is also available concerning the impact of cyanotoxins on other organisms, including humans. as a set of systemized data unique in the balkan peninsula, the database represents a possible model for other counno nco mm er cia l u se on ly z. svirčev et al.158 tries in the region and beyond. such databases encopassing all previous research (including monitoring and case reports), as well as continuous supplementation with the new available data are valuable in order to provide a timely and adequate reaction to toxic and noxious cyanobacteria, and thus prevent potential negative consequences. acknowlegdments the authors would like to acknowledge the funding from the ministry of education, science and technological development of the serbian government (project number: 176020) and cost action es1105 ‘cyanocost cyanobacterial blooms and toxins in water resources: occurrence, impacts and management’ for adding value to this study through networking and knowledge-sharing with european experts in the field. the authors wish to thank prof. jelena blaženčić for inspiration and scientific advice. references blaženčić j, 1986. [review of development of algology in serbia from 1883 to 1983].[article in serbian]. glasnik instituta za botaniku i botaničke bašte univerziteta u beogradu 20:99-108. blaženčić j, martinović-vitanović v, cvijan m, filipi-matutinović s, 1985. [bibliografija radova o algama i algološkim istraživanjima u sr srbiji od 1947. do 1980. godine].[article in serbian]. glasnik instituta za botaniku i botaničke bašte univerziteta u beogradu 19:233-266. chen j, dai j, zhang h, wang c, zhou g, han z, liu z, 2010. bioaccumulation of microcystin and its oxidative stress in the apple (malus pumila). ecotoxicology 19:796-803. ćirić m, marković z, dulić z, subakov-simić g, 2010. first report of cyanobacterium cylindrospermopsis raciborskii from carp ponds in serbia, p. 14. in: proceedings 8th int. conf. on toxic cyanobacteria (ictc8), istanbul, turkey. codd ga, metcalf js, beattie ka, 1999. retention of microcystis aeruginosa and microcystin by salad lettuce (lactuca sativa) after spray irrigation with water containing cyanobacteria. toxicon 37:1181-1185. crush je, briggs lr, sprosen jm, nichols sn, 2008. effect of irrigation with lake water containing microcystins on microcystin content and growth of ryegrass, clover, rape, and lettuce. environ. toxicol. 23:246-252. cvijan m, fužinato s, 2011. the first finding of cylindrospermopsis raciborskii (woloszińska) seenayya et subba raju 1972 (cyanoprokaryota) in serbia. arch. biol. sci. 63:507-510. cvijan m, fužinato s, 2012. cylindrospermopsis raciborskii (cyanoprokaryota)-potential invasive and toxic species in serbia. botanica serbica 36:3-8. drobac d, 2015. [putevi izloženosti čoveka cijanotoksinima i njihov uticaj na zdravlje].[phd thesis in serbian], university of novi sad, serbia. drobac d, svirčev z, tokodi n, vidović m, baltić v, božićkrstić v, lazić d, pavlica t, 2011. microcystins potential risk factors in carcinogenesis of primary liver cancer in serbia. geographica pannonica 13:70-80. drobac d, tokodi n, lujić j, marinović z, subakov-simić g, dulić t, važić t, nybom s, meriluoto j, codd ga, svirčev z, 2016. cyanobacteria and cyanotoxins in fishponds and their effects on fish tissue. harmful algae 55:66-76. đukić n, pujin v, maletin s, gajin s, gantar m, petrović o, ratajac r, seleši đ, matavulj m, 1991a. lentic waters eutrophication in vojvodina part i “borkovac”. zbornik radova instituta za biologiju 31:4-6. đukić n, pujin v, maletin s, gajin s, gantar m, petrović o, ratajac r, seleši đ, matavulj m, 1991b. lentic waters eutrophication in vojvodina part i “zobnatica”. zbornik radova instituta za biologiju 31:39-40. fužinato s, fodora a, subakov-simić g, 2010. arthrospira fusiformis (voronichina) komarek et lund (cyanoprokaryota) a new species for europe. algological studies 134:17-24. institute of public health of serbia, 2013. [monitoring kvaliteta vode jezera palić i ludaš i potoka kereš u 2012. godini].[report in serbian]. institute of public health of serbia, belgrade. jakovljević s, stanković s, 1931-1932. [particularites limnologiques des eaux kartiques de la region de belgrad].[article in french]. glasnik botaničkog zavoda i bašte univerziteta u beogradu 2:1-19. karadžić v, 2011. [eutrofikacija i njene posledice na primeru reke ponjavice (opština pančevo)].[phd thesis in serbian]. university of belgrade, serbia. karadžić v, subakov-simić g, krizmanić j, natić d, 2010. phytoplankton and eutrophication development in the water supply reservoirs garaši and bukulja (serbia). desalination 255:91-96. karadžić v, subakov-simić g, natić d, ržaničanin a, ćirić m, gačić z, 2013. changes in the phytoplankton community and dominance of cylindrospermopsis raciborskii (wolosz.) subba raju in a temperate lowland river (ponjavica, serbia). hydrobiologia 711:43-60. maslać m, obušković lj, jakovčev d, cakić p, tucović v, 1992. [previous investigations of opovo channel, one in the system of channels of pančevo, p. 28-33].[article in serbian]. in: konferencija o aktuelnim problemima zaštite voda “zaštita voda 92”, subotica. miljković d, vučković m, gotović d, milenković p, zarkov n, roški đ, 2004. water supply problems of majdanpek, p. 575-579].[article in serbian]. in: konferencija o aktuelnim problemima zaštite voda “zaštita voda 04“, borsko jezero. milovanović d, 1949. [bibliografski pregled algoloških istraživanja u srbiji do 1947. godine].[article in serbian]. glasnik prirodnjačkog muzeja u beogradu serija b: biološke nauke b 1-2:323-329. milovanović d, 1963. phytoplankton and primary production in fish ponds “koluta”. arch. biol. sci. 6:3-16. milovanović d, 1970. limnotypological changes of some waters as a consequence of meliorative works in river dunav hydrosystem near apatin. ekologija 5:55-70. milovanović d, 1973. phytoplankton structural changes in the first years of đerdap reservoir existence. arch. biol. sci. 25:75-83. milovanović d, živković a, 1953. plankton production investigations in ečka fish ponds. proceedings san 29:197-264. no nco mm er cia l u se on ly research on cyanobacteria in serbia 159 milovanović d, živković, a, 1959. phytoplankton production in živača fishpond (ii contribution to limnology of lentic waters in panonia valley). arch. biol. sci. 2:1-17. natić d, jovanović d, knežević t, karadžić v, bulat z, matović v, 2012. microcystin-lr in surface water of ponjavica river. vojnosanit. pregl. 69:753-758. obušković lj, 1982. phytoplankton and saprobiological characteristics of rivers bosut, spačva i studva. vodoprivreda 14:247-249. obušković lj, 1983. [das phytoplankton des stausees “eisernes tor” (đerdap) im jahre 1973].[article in german]. hidrobiologia 17:341-347. obušković lj, 1987. [phytoplankton and saprobiological characteristics of river sava in 1984, p. 426-430].[article in serbian]. in: proceedings conf. river sava, protection and water use, zagreb. obušković lj, 1989. [phytoplankton and saprobiological characteristics of river danube in 1988, p. 30-36].[article in serbian]. in: proceedings konferencija o aktuelnim problemima zaštite voda “zaštita voda 89”, rovinj. obušković lj, 1991. [phytoplankton and saprobiological characteristics of river ponjavice (south banat) as indicator of increased eutrophication, p. 332-337].[article in serbian]. in: proceedings konferencija o aktuelnim problemima zaštite voda “zaštita voda 91”, neum. perendija b, despotović s, radovanović t, gavrić j, borkovićmitić s, pavlović s, ognjanović b, simić s, pajović s, saičić z, 2011. biochemical and ultrastructural changes in the liver of european perch (perca fluviatilis l.) in response to cyanobacterial bloom in the gruža reservoir. arch. biol. sci. 63:979-989. protić đ, 1935. [hidrobiološke studije na kanalu kralja petra i kanalu kralja aleksandra. drugi deo].[article in serbian]. spomenik srpske kraljevske akademije 80:1-35. protić đ, 1936. [hidrobiološke studije na kanalu kralja petra. treći deo].[article in serbian]. spomenik srpske kraljevske akademije 85:59-87. pujin v, ratajac v, đukić n, svirčev z, kilibarda p, 1987. [saisonmassige variationen des zusammemensetzung des planktons und der bodenbesiedlung in der carska bara (jugoslawien) ].[article in german]. tiscia (szeged) 22:83-91. ristić o, gajin s, gantar m, matavulj m, 1979. [microbiological studies of some fish ponds in vojvodina, p. 19231935].[article in serbian]. in: proceedings ii congr. ecologists in yugoslavia, zagreb. saqurane s, ouahid y, el ghazali i, oudra b, bouarab l, del campo f, 2009. physiological changes in triticum durum, zea mays, pisum sativum and lens esculenta cultivars, caused by irrigation with water contaminated with microcystins: a laboratory experimental approach. toxicon 53:786-796. scharschmidt g, 1883. fragmenta phycologiae bosniaco serbicae].[article in german]. magyar novenytani lapok 75: 33-39. sedmak b, svirčev z, 2011. [cijanobakterije i njihovi toksiniekološki i toksikološki rizici i cvetanje cijanobakterija u srbiji].[book in slovenian]. visoka šola za varstvo okolja, velenje, slovenia: 134 pp. seleši đ, 1982. limnological investigations of lake ludoš. vode vojvodine 10:345-368. shllaku l, landner l, 1992. environment in kosovo: environmental problems related to mineral exploitation. stockholm, sweden, who. simeunović j, 2009. [ekofiziološke karakteristike potencijalno toksičnih i toksičnih vodenih sojeva cijanobakterija na području vojvodine].[phd thesis in serbian], university of novi sad, serbia. simić s, mišćević s, đorđević n, popović n, 2011. cyanobacteria in aleksandrovac lake before and after revitalisation. in: proceedings 16th conf. cyanobacteria and human health, academy of studenica, novi sad, serbia. subakov-simić g, plemić n, karadžić v, cvijan m, krizmanić j, 2004. [qualitative and quantitative analysis of phytoplankton in slatina near opovo, p. 327-330].[article in serbian]. in: proceedings konferencija o aktuelnim problemima zaštite voda “zaštita voda 04”, borsko jezero. svirčev z, simeunović j, subakov-simić g, krstić s, vidović, m, 2007. freshwater cyanobacterial blooms and cyanotoxin production in serbia in the past 25 years. geographica pannonica 11:12-21. svirčev z, krstić s, miladinov-mikov m, baltić v, vidović m, 2009. freshwater cyanobacterial blooms and primary liver cancer epidemiological studies in serbia. j. environ. sci. heal. c 27:36-55. svirčev z, baltić v, gantar m, juković m, stojanović d, baltić m, 2010. molecular aspects of microcystin-induced hepatotoxicity and hepatocarcinogenesis. j. environ. sci. heal. c 28:39-59. svirčev z, simeunović j, subakov-simić g, krstić s, pantelić d, dulić t, 2013a. cyanobacterial blooms and their toxicity in vojvodina lakes, serbia. int. j. environ. heal. r. 7:745-758. svirčev z, drobac d, tokodi n, vidović m, simeunović j, miladinov-mikov m, baltić v, 2013b. epidemiology of primary liver cancer in serbia and possible connection with cyanobacterial blooms. j. environ. sci. heal. c 31:181-200. svirčev z, tokodi n, drobac d, codd ga, 2014a. cyanobacteria in aquatic ecosystems in serbia: effects on water quality, human health and biodiversity. syst. biodivers. 12: 261-270. svirčev z, drobac d, tokodi n, lužanin z, munjas am, nikolin b, vuleta d, meriluoto j, 2014b. epidemiology of cancers in serbia and possible connection with cyanobacterial blooms. j. environ. sci. heal. c 32:319-337. svirčev z, obradović v, codd ga, marjanović p, spoof l, drobac d, tokodi n, petković a, nenin t, simeunović j, važić t, meriluoto j, 2016a. massive fish mortality and cylindrospermopsis raciborskii bloom in aleksandrovac lake. ecotoxicology 25:1353-1363. svirčev z, drobac d, tokodi n, đenić d, simeunović j, hiskia a, kaloudis t, mijović b, šušak s, protić m, vidović m, onjia a, nybom s, važić t, palanački malešević t, dulić t, pantelić d, vukašinović m, meriluoto j, 2016b. lessons from the užice case: how to complement analytical data. in: j. meriluoto, l. spoof and g.a. codd (eds.), handbook of cyanobacterial monitoring and cyanotoxin analysis. j. wiley & sons, chichester: 576 pp. tokodi n, drobac d, simeunović j, svirčev z, 2013. assessment of acute cyanotoxicity using artemia salina bioassay in water samples from fishponds. in: 17th international ecoconference, 10th environmental protection of urban and suburban settlements. 25-28 september 2013, novi sad, serbia. no nco mm er cia l u se on ly z. svirčev et al.160 tokodi n, drobac d, simeunović j, svirčev z. 2014. microcystin concentrations in fishpond waters. matice srpska j. nat. sci. 127:35-42. tokodi n, 2016. [toksične cijanobakterije sa teritorije republike srbije].[phd thesis in serbian], university of novi sad, serbia. who, 1999. toxic cyanobacteria in water: a guide to their public health consequences, monitoring, and management. world health organization, geneva, switzerland. world bank report, 2011. introducing daphnia grazing to control global warming associated cyanobacterial toxic blooms in fishing pond. report dm 4307. in: world health organization 1998. guidelines for drinking-water quality, 2nd ed. addendum to vol. 2. who, geneva, switzerland. no nco mm er cia l u se on ly layout 1 introduction corbicula fluminea (müller, 1774) (mollusca: bivalvia: veneroidea: cyrenidae) is considered among the 100 worst invasive species in the world (lowe et al., 2000). it has spread over the last century in europe and north and south america, endangering the biodiversity of most aquatic ecosystems in these continents and progressively increasing its range (mcmahon, 1982; araujo et al., 1993; marescaux et al., 2010). this bivalve has a very complex life history, including clonal and androgenetic reproduction, which has complicated the interpretation by molecular analyses of invasive populations (renard et al., 2000; hedtke et al., 2008; pigneur et al., 2011). in addition, no useful monographs have been published dealing with the different taxa in this genus in their native area, which added to the apparent intraspecific variability of c. fluminea, further complicates the study of invasive populations of corbicula clams. apart from c. fluminea, a number of other taxa have also been characterized as invasive in europe and the americas. the best-known species is corbicula fluminalis (müller, 1774), a taxon long considered either a synonym or a morphotype of c. fluminea. however, different studies have demonstrated that it is a valid species, both by molecular and anatomical characteristics (renard et al., 2000; glaubrecht et al., 2003; mouthon and parghentanian, 2004; hedkte et al., 2008). it is present in no less than ten european countries, including italy, spain, france, switzerland, belgium, netherlands, germany, poland, hungary and serbia (csányi, 1999; chevallier, 2003; korniushin, 2004; mouthon and parghentanian, 2004; cianfanelli et al., 2007; ciutti and cappelletti, 2009; bódis et al., 2011; quiñonero salgado and lópez soriano, 2014), and in south america (martins et al., 2006; pereira et al., 2014). less known is corbicula largillierti (philippi, 1844), a species widely distributed in south america, where it usually lives syntopically with c. fluminea (ituarte, 1994; martins et al., 2006; reyna et al., 2013; torre and reyna, 2013; azevedo et al., 2014; pereira et al., 2014), but also recently reported from europe (quiñonero salgado and lópez soriano, 2016a; nesemann, 2018). genetic studies suggest that c. fluminea could be mixed with other taxa in some populations (pigneur et al., 2011), and different invasion episodes in some places have been proposed (chevallier, 2003; peñarrubia et al., 2016). recently, growing evidence suggests that invasive populations of corbicula may hide cryptic or uncharacterized taxa. for example in north america, after more than 80 years with only a single described species, c. fluminea, a new study revealed at least three taxa, with distinct morphology and molecular markers, live syntopically in the illinois river (tiemann et al., 2017). in south america, c. fluminalis and an unknown species have also been described in addition to c. fluminea and c. largillierti (clavijo, 2014; pereira et al., 2014). the most complex situation has been described in western europe, with up to seven distinctive taxa inhabiting advances in oceanography and limnology, 2018; 9(2): 97-104 article doi: 10.4081/aiol.2018.7857 this work is licensed under a creative commons attribution-noncommercial 4.0 international license (cc by-nc 4.0). unraveling the complexity of corbicula clams invasion in lake garda (italy) joaquín lópez-soriano,1* sergio quiñonero-salgado,1 cristina cappelletti,2 filippo faccenda,2 francesca ciutti2 1associació catalana de malacologia-museu blau, plaça leonardo da vinci 4-5, barcelona, spain; 2technology transfer centre, fondazione edmund mach, san michele all'adige, italy abstract lake garda, the largest italian lake, is suffering from the introduction of several non-indigenous species during the last decades and can now be considered one of the main european freshwater hotspots of xenodiversity. among the bivalvia (veneroidea, cyrenidae), corbicula fluminea and corbicula fluminalis were first recorded in 2002 and 2008 respectively, and are now widespread in the southern part of the basin. recent observation of specimens that did not resemble either of these taxa, suggested that the populations of invasive corbicula of lake garda could include some other taxa not previously recorded. with this aim, a thoroughly characterization of corbicula shells found at lake garda was made. by studying morphometric parameters and comparing them with specimens collected in spain (ebro and ter rivers), the presence in lake garda of two other related species not previously recorded in italy, namely c. leana and c. largillierti, has been confirmed. the syntopic presence of at least four species of this genus in a single environment is a singular occurrence both in italy and europe. key words: corbicula leana; corbicula largillierti; cryptic species; morphometrics; invasions; xenodiversity hotspot. received: october 2018. accepted: december 2018. no nco mm er cia l u se on ly j. lópez-soriano et al.98 aquatic systems in ne iberian peninsula, all characterized by distinctive morphological characters and juveniles, with no apparent intermediate forms or hybridization (quiñonero salgado and lópez soriano, 2016a; 2016b). one of these taxa has been characterized as corbicula leana prime, 1867, whose presence was previously postulated by molecular markers in france (pigneur et al., 2011), and later confirmed by morphologic traits both in spain and france (prié, 2017; quiñonero salgado and lópez soriano, 2017). c. largillierti has also been reported for the first time in europe, in spain (quiñonero salgado and lópez soriano, 2016a) and later in germany (nesemann, 2018), in populations where only c. fluminea had been described before. in lake garda, the asian clam c. fluminea was first recorded in 2002 in a very localized littoral area, and further studies backdated its introduction to 2000 (nardi and braccia, 2004; ciutti et al., 2007). moreover c. fluminalis was found in july 2008 (ciutti and cappelletti, 2009). the observation of specimens found in the last years, which clearly do not resemble any of the previously described species in lake garda and in italy, suggest that they could include some other not yet described taxa, and so the populations of invasive corbicula should be thoroughly characterized in order to better know their composition. with this aim, we made a preliminary study on the morphometric distinctiveness of different corbicula populations collected from the lake, and compared them with other european populations where similar morphs had been previously described. methods lake garda, the largest italian lake (volume=49 km3, area=368 km2 and maximum depth=350 m) is one of the deep insubrian lakes south of the alps, along with lakes iseo, como, lugano and maggiore. the lake is included in the ilter network (italian long term ecological research) and its trophic condition ranges from oligotrophy to mesotrophy (salmaso, 2005; salmaso et al., 2009). shells of corbicula were collected from lake garda between march and october 2017 in sampling stations along the lake shore and sediment samples by using a grab (maximum depth of 30 m) (fig. 1). shells of different corbicula species collected by the authors in ebro and ter rivers (ne spain) were used for comparison. for the morphological analysis, a total of 879 collected shells were used (438 from italy and 441 from spain), divided into 4 size categories according to shell length in order to eliminate bias caused by allometric growth: 6-9 mm, 10-20 mm, 21-30 mm and 31-40 mm (bódis et al., 2011; reyna et al., 2013). the following morphometric variables were measured to the nearest 0.1 mm with digital caliper: shell length (l) (antero-posterior dimension) and shell height (h) (distance from the umbo to the ventral margin). the shell mass (w) was weighted by analytical scale (accuracy 0.001 g). a principal component analysis (pca) was carried out using the morphometric parameters shell length (l), shell height (h) and weight (w) (correlation matrix) for specimens belonging to the shell length class range 10-20 mm, common to all studied populations (n=626). the value of the morphometric variables did not reveal a normal distribution according to the shapiro-wilk test and this justified the use of nonparametric statistics. the significance of the differences between the various morphometric parameters of the individuals was evaluated using a rank-based nonparametric anova kruskal-wallis and post hoc comparison (z test). all statistical analysis were performed with the statistica 13.1 software (dell inc., tulsa, ok, usa), setting the significant level of p≤0.05. fig. 1. sampling stations in lake garda. no nco mm er cia l u se on ly corbicula invasion in lake garda 99 results based on morphological features, two new morphs of corbicula clams found in lake garda, living syntopically with c. fluminea and c. fluminalis, were determined as c. largillierti and c. leana, showing similarities with spanish populations (quiñonero salgado and lópez soriano, 2016a, 2016b). corbicula specimens collected in lake garda are shown in fig. 2. morphometric data of c. fluminea, c. fluminalis, c. largillierti and c. leana from spain and lake garda showed a high variability in shell dimensions (tab. 1). in particular, c. fluminea and c. leana specimens from spain were longer and heavier, with maximum shell length (l) up to 40 mm and maximum weight of 13.7 g and 12.8 g respectively. an evident separation of c. fluminalis and c. leana for both lake garda and spanish populations was observed with pca, with the two principal components explaining most variation (99.5%); c. fluminea and c. largillierti were less clearly separated (fig. 3). significant differences fig. 2. specimens of corbicula from lake garda. c. fluminea (scale: a, 0.5 cm; b-d, 1 cm); c. fluminalis (scale: 1 cm); c. leana (scale: a, 0.5 cm; b-d, 1 cm). c. largillierti (scale: a, 0.5 cm; b-d, 1 cm). no nco mm er cia l u se on ly j. lópez-soriano et al.100 for all considered variables (l, h, w, h/l and l/w) were observed for two length classes: 10-20 mm and 21-30 mm (kruskal-wallis test) (tab. 2). in the 10-20 mm class (common to all populations from lake garda and spain) no significant differences were observed between lake garda and spain after post hoc comparison for c. leana and c. largillierti for all variables. also within the 21-30 mm class, c. leana from lake garda was not different from the spanish population. significant differences were observed for c. fluminea and c. fluminalis in the size range 10-20 mm and for c. fluminea in the size range 21-30 mm. a comparison within the 31-40 mm class was not performed because it included just two populations (c. leana and c. fluminea from spain). the height/length ratio (h/l) and the length/weight (l/w) ratio can be descriptive of the differences in the shape and weight of the species and the l/w ratio in particular seemed to be relatively constant for each taxon (fig. 4). in the shell length 10-20 mm class, c. fluminalis from lake garda and spain had significantly a higher height/length ratio (h/l), with mean values of 1.15 and 1.07 respectively, corresponding to a less elongated form; on the contrary, c. leana had the lowest values (0.88 and 0.80 respectively). the h/l distinction among other species was less clear. the post hoc comparison also indicated significant differences between c. fluminea from lake garda and spain; differences were also observed between c. fluminalis (italy and spain). no differences between lake garda and spanish populations were found for c. largillierti and c. leana. within the shell length 21-30 mm class the h/l ratio showed no significant differences between lake garda and spain for c. fluminea. tab. 1. morphometric data of corbicula shells from lake garda (italy) and spain. h (mm) l (mm) w (g) n mean sd range mean sd range mean sd range c. fluminea g 225 17.0 4.0 6.4-28.6 17.9 4.1 7.7-30.1 1.705 0.948 0.068-5.737 c. fluminalis g 104 13.5 2.6 8.0-21.8 11.8 2.6 7.6-20.0 1.225 0.802 0.213-4.385 c. largillierti g 60 14.0 2.1 8.0-18.2 14.9 2.2 8.8-19.2 0.704 0.291 0.112-1.361 c. leana g 49 12.4 2.9 7.4-19.8 14.1 3.0 9.1-22.3 0.674 0.474 0.110-2.303 c. fluminea s 130 22.5 9.6 7.0-37.8 24.5 9.7 8.2-40.4 3.921 3.294 0.124-13.700 c. fluminalis s 168 18.2 3.8 6.0-25.3 17.2 3.7 5.9-24.3 2.756 1.312 0.094-6.357 c. largillierti s 20 15.0 1.9 11.4-18.2 15.9 1.7 13.0-18.9 0.798 0.307 0.320-1.360 c. leana s 123 20.0 8.4 7.5-37.0 23.3 8.8 9.8-40.8 3.034 2.822 0.145-12.841 g, lake garda, italy; s, spain. fig. 3. the principal component analysis showing the relationship of the pc1 and pc2 components of corbicula populations from lake garda (italy) and spain (10-20 mm). red, c. fluminalis; black, c. fluminea; green, c. largillierti; blue, c. leana. filled circle, lake garda; empty circle, spain. no nco mm er cia l u se on ly corbicula invasion in lake garda 101 considering the shell length range common to all populations (10-20 mm), the l/w ratio was more informative to distinguish species. c. fluminalis from lake garda and spain showed significantly lower values (9.96 and 7.98 respectively) in comparison to other taxa, corresponding to heavier shells. c. leana and c. largillierti had significantly higher values respect to c. fluminea. for the l/w ratio, no significant differences for all species from lake garda and spain were observed. discussion with the new observation of c. leana and c. largillierti reported in this paper, lake garda corbicula populations include at least four species, instead of the two previously cited (c. fluminea and c. fluminalis). indeed, we confirm the presence of two species not previously cited in italy, which expands to a third european country the distribution of both c. leana and c. largillierti. besides morphological features (ribs, colour of teeth and inner and outer shell), morphometric parameters (h/l and l/w ratios in particular) are informative to distinguish among the four analysed species (quiñonero salgado and lópez soriano, 2016a, 2016b) (tab. 3). c. leana is characterized by a very uniform orange umbo and violet lateral teeth, as described in quiñonero salgado and lópez soriano (2016a; 2017); it has been described as having a shell that is consistently broader and lighter than c. fluminea, as also shown in the morphometric analysis. on the other hand, c. largillierti is not so clearly discernible by morphometric parameters but has peculiar tab. 2. results of the anova kruskal-wallis for the morphometry of corbicula shells from lake garda and spain grouped in three length classes: 6-9 mm, 10-20 mm, 21-30 mm lenght class 6-9 mm 10-20 mm 21-30 mm df=6; n=20 df=7; n=626 df=4; n=127 h p h p h p h (mm) 4,5102 ns 195,2877 *** 31,6538 *** l (mm) 4,7255 ns 125,4989 *** 66,7735 *** w (g) 7,3996 ns 279,6785 *** 64,1655 *** h/l 16,8168 ns 548,8888 *** 101,2741 *** l/w 11,8238 ns 386,3458 *** 70,8472 *** significance level p<0.05;(ns, not significative; ***p<0.001. fig. 4. boxplot (median and 25-75% quartiles) of a) height/length (h/l) ratio and b) length/weight (l/w) ratio for specimens in the range 10-20 mm of corbicula specimens. flu, c. fluminea; fls, c. fluminalis; lar, c. largillierti; lea, c. leana; g, lake garda; s, spain. no nco mm er cia l u se on ly j. lópez-soriano et al.102 morphological features: a dark violet coloration on the entire inner surface, and lacks the blue blotch in the umbo that is always present in c. fluminea (quiñonero-salgado and lópez soriano, 2016a, 2016b). c. largillierti and c. leana do not have significant differences in their morphometric parameters and ratios between lake garda and spanish populations. not surprisingly, the most studied and widespread species, c. fluminea and c. fluminalis, show significant biometric differences between lake garda and spain, as already observed in other countries (araujo et al., 1993; sousa et al., 2007; skuza et al., 2009; gomes et al., 2016). the syntopic presence of up to four species in a single environment is rather a new scenario, but it could indeed be a more widespread situation than expected, since no in-depth analyses have been performed in most of the corbicula invasive populations. it has to be considered that the number of alien aquatic species in lake garda, -up to 42 species of fish, invertebrates, macrophytes and macroalgaeis quite high and constantly growing, probably due to tourism and recreational activities such as fishing and sailing (ciutti et al., 2011; ciutti and cappelletti, 2017). indeed, the abundance of these invasive species has shown dramatic increases, as seen in accumulation of shells in the shore of the lake. only a few years after their introduction, corbicula clams are now widespread in the southern part of the lake and large deposits along the shore can be observed, resulting in an impressive scenario (cappelletti and ciutti, 2017) (fig. 5). we cannot speculate on the likely origin and time of arrival for these two newcomer species. however, no large specimens of c. leana were found in lake garda, compared to those found in the ter river in spain, which may reach up to 40 mm when adult. thus, a likely recent origin could be postulated for this population in the lake. in this sense, this species was not detected in this lake before 2017. instead some specimens of corbicula sp., now assigned to c. largillierti, were already found in march 2015 (shell length up to 16.7 mm). tab. 3. main morphological features of c. fluminea, c. fluminalis, c. largillierti and c. leana (according to quiñonero salgado and lópez soriano, 2016a, 2016b). morphometric data from this study (mean and standard deviation). c. fluminea c. fluminalis c. largillierti c. leana shell outer color olive green/olive/brown brown brown yellow/brown shell inner color ivory/light violet purple/dark violet purple/dark violet ivory/orange shell outline oval to triangular triangular oval to triangular oval, wide/lengthened lateral teeth violet blotch on base no blotches no blotches stained violet umbo olive green/brown; brown violet, without ribs orange vertical violet line barely visible ribs coarse, sparse regular, thick regular, thick coarse, sparse juveniles green, violet line in umbo purple/violet, no line in umbo purple/violet, wide stripe in umbo orange, no line in umbo h/l lake garda (i) 0.95 (0.04) 1.15 (0.06) 0.94 (0.03) 0.88 (0.03) h/l ebro/ter (s) 0.87 (0.03) 1.07 (0.04) 0.94 (0.03) 0.80 (0.03) l/w lake garda (i) 14.00 (5.33) 9.96 (2.78) 23.39 (6.62) 28.65 (14.38) h/l ebro/ter (s) 16.65 (5.65) 7.98 (3.94) 22.62 (7.84) 27.32 (14.73) fig. 5. large deposits of corbicula shells along the lake garda shore.no nco mm er cia l u se on ly corbicula invasion in lake garda 103 the new analysis of corbicula populations in lake garda confirms that some populations may hide other corbicula species apart from c. fluminea and c. fluminalis, as already suggested for different sites in north and south america, france, germany and spain (torre and reyna, 2013; quiñonero salgado and lópez-soriano, 2016a; tiemann et al., 2017; nesemann, 2018). indeed, one of the species present in the ebro river has been found in public and private collections, sampled around the year 2000 (quiñonero salgado and lópez soriano, 2016a), thus confirming a cryptic presence for many years, misidentified as c. fluminea. in fact, the identification of c. fluminea has not followed a true “gold standard” criterion in the last decades, which is also evident because most of the published studies do not even include pictures of the collected specimens. furthermore, in many cases citations were only based on few, bad preserved and dead specimens, which may not maintain some important diagnostic characteristics. in the light of our results, and those by other authors, a detailed revision of the invasive populations of corbicula is advisable, as some populations may have a richer species composition than reported. indeed, our results confirm the predictions made by reyna et al. (2018), based on bioclimatic models and known distribution ranges of corbicula species, who postulated that c. largillierti could potentially expand its invasive range to different countries in europe, with likely sympatric occurrence in some places with c. fluminea and c. fluminalis. increase in global trade and tourism may favor in a near future these episodes of colonization. conclusions with these results, we increase the number of corbicula species present in italy to four, and describe the first known populations of c. leana and c. largillerti in this country. for europe, c. leana is present in spain and france, while c. largillerti has only been reported in spain and germany. other italian populations should be studied, in order to detect the expansion of the four species already found in lake garda. acknowledgments we warmly acknowledge federica giacomazzi (agenzia regionale per la prevenzione e protezione ambientale veneto arpav, dipartimento regionale laboratori, servizio laboratorio di verona) for providing corbicula specimens from lake garda and dr. giovanna flaim for english revision. also thanks to álvaro alonso for assistance with the corbicula plate. references araujo r, moreno d, ramos ra, 1993. the asiatic clam corbicula fluminea (müller, 1774) (bivalvia: corbiculidae) in europe. amer. malac. bull. 10:39-49. azevedo el, barbosa jel, vidigal thd, callisto m, molozzi j, 2014. first record of corbicula largillierti (philippi, 1844) in the paraíba river basin and potential implications from water diversion in the sao francisco river. biota neotrop. 14:1-4. bódis e, nosek j, oertel n, tóth b, fehér z, 2011. a comparative study of two corbicula morphs (bivalvia, corbiculidae) inhabiting river danube. int. rev. hydrobiol. 96:257-273. cappelletti c, ciutti f, 2017. [bivalvi alloctoni nel lago di garda].[article in italian]. biologia ambientale 31:169-173. chevallier h, 2003. nouvelles données sur la propagation en france des corbicula mergerl von mühlfeld, 1811 (bivalvia: corbiculidae). doc. malacol. 4:11-14. cianfanelli s, lori e, bodon m, 2007. non-indigenous freshwater molluscs and their distribution in italy, p. 103-121. in: f. gherardi (ed.), biological invaders in inland waters: profiles, distribution, and threats. springer, dordrecht. ciutti f, girod a, mariani m, 2007. [considerazioni su una popolazione di corbicula fluminea (müller, 1774) nel lago di garda sud-orientale (italia)].[article in italian]. ann. mus. civ. sc. nat. brescia. 35:121-124. ciutti f, cappelletti c, 2009. first record of corbicula fluminalis (müller, 1774) in lake garda (italy), living in sympatry with corbicula fluminea (müller, 1774). j. limnol. 68:162-165. ciutti f, beltrami me, confortini i, cianfabelli s, cappelletti c, 2011. non-indigenous invertebrates, fish and macrophytes in lake garda (italy). j. limnol. 70:315-320. ciutti f, cappelletti c, 2017. [invasioni biologiche: il caso del lago di garda].[article in italian]. biologia ambientale 31: 59-164. clavijo c, 2014. [diversidad de corbiculidae (mollusca: bivalvia) en uruguay].[ph.d. thesis in spanish], universidad de la república, montevideo. csányi b, 1999. spreading invaders along the danubian highway: first record of corbicula fluminea (o.f. müller, 1774) and c. fluminalis (o.f. müller, 1774) in hungary (mollusca: bivalvia). fol. hist. nat. mus. matr. 23:343-345. glaubrecht m, von rintelen t, korniushin av, 2003. toward a systematic revision of brooding freshwater corbiculidae in southeast asia (bivalvia, veneroida): on shell morphology, anatomy and molecular phylogenetics of endemic taxa from islands in indonesia. malacologia 45:1-40. gomes c, sousa r, mendes t, borges r, vilares p, vasconcelos v, et al. 2016. low genetic diversity and high invasion success of corbicula fluminea (bivalvia, corbiculidae) (müller, 1774) in portugal. plos one 11:e0158108. hedtke sm, stanger-hall k, baker rj, hillis dm, 2008. allmale sexuality: origin and maintenance of androgenesis in the asian clam corbicula. evolution 62:1119-1136. ituarte c, 1994. corbicula and neocorbicula (bivalvia: corbiculidae) in the paraná, uruguay, and río de la plata basins. nautilus 107:129-135. korniushin av, 2004. a revision of some asian and african freshwater clams assigned to corbicula fluminalis (müller, no nco mm er cia l u se on ly j. lópez-soriano et al.104 1774) (mollusca: bivalvia: corbiculidae), with review of anatomical characters and reproductive features based on museum collections. hydrobiologia 529:251-270. lowe s, browne m, boudjelas s, de poorter m, 2000. 100 of the world’s worst invasive alien species. a selection from the global invasive species database (iucn). the invasive species specialist group: 12 pp. marescaux j, pigneur lm, van dononck k, 2010. new records of corbicula in french rivers. aquat. invasions 5:s35-s38. martins ds, veitenheimer-mendes il, faccioni-heuser mc, 2006. morphological aspects and incubation in three species of corbicula mühlfeld, in guaíba lake, rio grande do sul, brazil (bivalvia, corbiculidae). biota neotrop. doi.org/ 10.1590/s1676-06032006000200016 mcmahon rf, 1982. the occurrence and spread of the introduced asiatic freshwater clam, corbicula fluminea (müller) in north america: 1924-1982. nautilus 96:134-141. mouthon j, parghentanian t, 2004. comparison of the life cycle and population dynamics of two corbicula species, c. fluminea and c. fluminalis (bivalvia: corbiculidae) in two french canals. arch. hydrobiol. 161:267-287. nardi g, braccia a, 2004. [prima segnalazione di corbicula fluminea (o.f. müller, 1774) per il lago di garda (provincia di brescia) (mollusca bivalvia, corbiculidae)].[article in italian]. boll. soc. malac. ital. 39:181-184. nesemann hf, 2018. [corbicula largillierti im oberrhein (hessen), neu erkannt in deutschland].[article in german]. mitt. dtsch. malakozool. ges. 98:65-68. peñarrubia l, viñas j, vidal o, sanz n, pla c. araguas rm, 2016. genetic characterization of the asian clam species complex (corbicula) invasion in the iberian peninsula. hydrobiologia 784:349-365. pereira d, mansur mcd, duarte lds, de oliveira as, piampão dm, callil ct, ituarte c, parada e, peredo s, darrigran g, scarabino f, clavijo c, lara g, miyahira ic, rodríguez mtr, lasso c, 2014. bivalve distribution in hydrographic regions in south america: historical overview and conservation. hydrobiologia 735:15-44. pigneur lm, marescaux j, roland k, etoundi e, descy jp. van doninck k, 2011. phylogeny and androgenesis in the invasive corbicula clams (bivalvia, corbiculidae) in western europe. bmc evol. biol. 11:147. prié v, 2017. [naïades et autres bivalves d’eau douce de france].[book in french]. muséum national d’histoire naturelle, biotope, mèze: 336 pp. quiñonero salgado s, lópez soriano j, 2014. [presencia de corbicula fluminalis (o.f. müller, 1774) (bivalvia: corbiculidae) en el bajo ebro (cataluña)].[article in spanish]. spira 5:139-141. quiñonero salgado s, lópez soriano j, 2016a. [el género corbicula mühlfeld, 1811 (corbiculidae: bivalvia) en el bajo ebro (ne de la península ibérica) ].[article in spanish]. nemus 6:9-33. quiñonero salgado s, lópez soriano j, 2016b. [revisión del género corbicula von mühlfeld, 1811 (bivalvia: corbiculidae) en el baix empordà (ne península ibérica)].[article in spanish]. spira 6:29-39. quiñonero salgado s, lópez soriano j, 2017. presence of corbicula leana prime, 1867 (bivalvia: corbiculidae) in france. avenionia 2:82-87. renard e, bachman v, cariou ml, moreteau jc, 2000. morphological and molecular differentiation of the invasive freshwater species of the genus corbicula (bivalvia, corbiculidea) suggest the presence of three taxa in french rivers. mol. ecol. 9:2009-2016. reyna pb, morán ag, tatián m, 2013. taxonomy, distribution and population structure of invasive corbiculidae (mollusca, bivalvia) in the suquía river basin, córdoba, argentina. iheringia ser. zool. 103:77-84. reyna p, nori j, ballesteros m, hued a, tatián m, 2018. targeting clams: insights into the invasive potential and current and future distribution of asian clams. environ. conserv. 45:387-395. salmaso n, 2005. effects of climatic fluctuations and vertical mixing on the interannual trophic variability of lake garda, italy. limnol. oceanogr. 50:553-565. salmaso n, boscaini a, cappelletti c, ciutti f, 2009. [le condizioni di salute del lago di garda: aggiornamento dello stato delle conoscenze su carichi di nutrienti algali e sulle componenti biologiche della zona pelagica e litorale], p. 4988 .[article in italian]. in: f. bertin and a. bortoli (eds.), proceedings national congressng “problematiche ambientali del lago di garda. approfondimenti e proposte di risanamento”], torri del benaco. skuza l, łabęcka am, domagała j, 2009. cytogeneticand morphological characterization of corbicula fluminalis (o.f. müller, 1774) (bivalvia: veneroida: corbiculidae): taxonomic status assessment of a freshwater clam. folia biol.-krakow 57:177-185. sousa r, freire r, rufino m, méndez j, gaspar m, antunes c, guilhermino l, 2007. genetic and shell morphological variability of the invasive bivalve corbicula fluminea (müller, 1774) in two portuguese estuaries. estuar. coast. shelf s. 74:166-174. tiemann js, haponski ae, douglass sa, lee t, cummings ks, davis ma, foighil dó, 2017. first record of a putative novel invasive corbicula lineage discovered in the illinois river, illinois, usa. bioinv. rec. 6:159-166. torre l, reyna p, 2013. bivalvia, veneroidea, corbiculidae, corbicula largillierti (philippi, 1844): new distribution record in the del valle central basin, catamarca province, argentina. check list 9:165-166. no nco mm er cia l u se on ly layout 1 introduction blooms of cyanobacteria (blue-green algae/ cyanoprokaryotes) have increased globally in recent decades (paerl and otten, 2013; harke et al., 2016). due to the ability of toxin production, some species affect livestocks and high cyanotoxin concentrations were linked to animal deaths and human health hazard through drinking and recreational waters (codd et al., 1999; carmichael et al., 2001; azevedo et al., 2002; backer et al., 2015). cyanobacteria can produce different types of toxic compounds, which include hepatotoxins, neurotoxins, cytotoxins, dermatotoxins and irritant toxins (bláha, 2009; westrick et al., 2010). the occurence of cyanotoxins have been reported in several cyanobacterial genera such as microcystis, nodularia, aphanizomenon, planktothrix, anabaena and cylindrospermopsis (sivonen et al., 1990; merel et. al., 2013; bernard et al., 2017). the most studied group of cyanobacterial toxins are the hepatotoxic cyclic peptides, which include the microcystins and nodularins. although they are similar in structure, nodularin has been isolated from only one species of cyanobacteria, nodularia spumigena mertens ex bornet & flahault, whereas microcystin can be produced by multiple cyanobacterial genera, most notably by microcystis, planktothrix or anabaena (sivonen and jones, 1999; bernard et al., 2017). over 100 microcystin variants and 10 nodularin variants have been identified (spoof et al., 2001; bortoli and volmer, 2014). cyanobacterial blooms occur in turkish inland waters, mostly lakes and reservoirs used as supplies of drinking water or recreation. aphanizomenon sp. was the first cyanobacteria to cause problems in filter system of drinking water treatment plant in kurtbogazi dam lake (ankara) in 1981 (guler aykulu, pers. comm.). during the 1990s many cyanobacterial blooms were detected in the marmara region. in 1994, blooms of anabaena spp. resulted in fish mortality in i̇znik lake (albay et al., 2003a). cyanotoxin research has started at the end of 1990s and increased in recent years (albay et al., 2003a,b; albay et al., 2005; akçaalan et al., 2006, 2014a, 2014b, 2016) it is well known that microscopic identification of cyanobacteria is time consuming and it requires taxonomic expertise. due to this limitation, molecular tools have been increasingly applied also to environmental studies (kurmayer and christiansen, 2009; bukowska et al., 2014). especially, because of the conserved nature of the 16s rrna gene, it is used to discriminate strains at the species level (neilan et al., 1997; moffitt and neilan, advances in oceanography and limnology, 2017; 8(1): 52-60 article doi: 10.4081/aiol.2017.6394 this work is licensed under a creative commons attribution-noncommercial 4.0 international license (cc by-nc 4.0). molecular detection of hepatotoxic cyanobacteria in inland water bodies of the marmara region, turkey latife köker,1 reyhan akçaalan,1 meriç albay,1 brett a. neilan2 1istanbul university, fisheries faculty, ordu cad. no:200 34470, laleli istanbul, turkey; 2school of biotechnology and biomolecular sciences, university of new south wales, 7 sydney 2052, australia *corresponding author: latifekoker@gmail.com abstract blooms of cyanobacteria are an increasingly frequent phenomenon in freshwater ecosystems worldwide as a result of eutrophication. many species can produce hepatotoxins that cause severe health hazards to humans. the aim of this study was to identify the bloom forming cyanobacteria species by molecular methods and to amplify genes responsible for hepatotoxin biosynthesis from the environmental samples and isolated strains of cyanobacteria from küçükçekmece lagoon, sapanca, i̇znik, manyas and taşkısı lakes. a total of 10 bloom samples and 11 isolated strains were examined and microcystis spp., planktothrix spp., nodularia spumigena, anabaenopsis elenkinii, sphaerospermopsis aphanizomenoides, cylindrospermopsis raciborskii were identified. hepatotoxin genes were detected in 60% of the bloom samples and 45% of the strains. two microcystis strains were obtained from küçükçekmece lagoon. while the strain assigned to microcystis flosaquae was non-toxic, microcystis aeruginosa strain produced microcystin. according to pcr results, the m. aeruginosa and planktothrix agardhii bloom samples of küçükçekmece lagoon contained the microcystin synthetase gene e (mcye) indicative of microcystin production, however, no microcystin was detected by hplc. the mcye gene was also found in microcystis wesenbergii isolated from taşkısı lake, and in all planktothrix rubescens bloom samples from sapanca lake. to our knowledge, this is the first detailed study for identifiying different toxic cyanobacteria species and their hepatotoxin production from several waterbodies in turkey using molecular methods. key words: 16s rrna, aminotransferase, nodularin, microcystin, cyanobacteria, cyanotoxin. received: november 2016. accepted: april 2017. no nco mm er cia l u se on ly molecular identification of hepatotoxic cyanobacteria in turkey 53 2001). jungblut and neilan (2006) developed a molecular method to detect both microcystin and nodularin-producing species by amplifying and sequencing of the aminotransferase (amt) domain of mcye and ndaf genes in the mcy and nda operons. the reason for choosing amt domain was its important role in synthesis of all microcystins and nodularins. due to the increased frequency of algal blooms in turkish lakes, it is important to understand the distribution of toxin-producing cyanobacteria in this area. the aims of the present study were to determine the bloom-forming cyanobacteria species using the 16s rrna gene as well as the potential toxicity using the mcye and ndaf genes indicative of microcystin/nodularin biosynthesis, occurring in lakes around the marmara region (küçükçekmece, sapanca, i̇znik, manyas and taşkısı). methods sampling sites cyanobacterial blooms have been collected from five lakes in marmara region (fig. 1). i̇znik lake, located in the southeast of marmara region, is the fifth biggest lake in turkey. cyanobacterial blooms occured because of heavy nutrient loading (albay et al., 2003a; akçaalan et al., 2006; tas and gonulol, 2007). the first bloom was formed by anabaena sp. in 1994. planktothrix rubescens (de candolle ex gomont) anagnostidis & komárek and nodularia spumigena were also detected (akçaalan et al., 2006; akçaalan et al., 2009). sapanca lake is an oligomesotrophic lake and planktothrix rubescens blooms have been observed in the metalimnion of the lake since the 1980s (akçaalan et al., 2006). the other studied area, küçükçekmece lagoon (istanbul, turkey), has a connection to the marmara sea via a narrow channel. the lagoon is in hypereutrophic conditions and microcystis aeruginosa (kützing) kützing blooms were observed from late spring to mid-autumn (albay et al., 2005) manyas lake is a eutrophic lake which is an important bird sanctuary, and in 1998 it was listed in the ramsar convention (çelik and ongun, 2006). taşkısı lake is a small, shallow lake situated in the eastern part of the marmara region (aykulu et al., 1999) (tab. 1). cyanobacteria identification freshly collected bloom samples were identified by inverted microscopy (axio observer z1, carl zeiss gmbh, jena, germany). 1-2 drops of fresh sample were fig. 1. location of sampling lakes in marmara region. no nco mm er cia l u se on ly l. köker et al.54 investigated according to taxonomical keys using filament/colony traits, presence and structure of mucilage, cell shape and size, whether having a specialized cell or not. cyanobacterial identification was done according to whitton and potts (2007), komárek (2013), komárek and anagnostidis (1986; 1999; 2005) and anagnostidis and komárek (1988). environmental samples during 2004-2009, ten bloom samples were collected from five lakes of marmara region (tab. 2). for cyanotoxin and molecular analysis, samples were collected using plankton net (20 µm mesh size, hydro-bios) and lyophilised and conserved at -20°c. cyanobacterial strains cyanobacterial strains used in the present study (tab. 3) were collected from blooms. single filaments and colonies of cyanobacteria were isolated by repeated washing with sterile media from a pasteur pipette and transferred 96-well plates filled with 200 µl bg 11 medium with or without nitrate according to presence or absence of heterocytes (rippka et al., 1979). dna extraction dna extraction from fresh cell pellets and lyophilized bloom samples was performed using xs extraction buffer containing 1% potassium-methylxanthogenate (800 mm ammonium acetate; 20 mm edta; 1% sds; 100 mm tris-hci, ph 7.4) (tillett and neilan, 2000). dna was dissolved in tris-edta buffer (10:1). concentrations of dna were determined using a nanodrop® nd-1000 spectrophotometer and dna extracts were stored at -20°c. pcr amplification and sequencing all pcr reactions were performed in 20 µl reaction volume containing pcr buffer (bioline, london, uk), 2.5 mm mgci2, 0.2 mm dntps (bioline), 10 pmol each of the forward and reverse primers and 0.2 u taq polymerase (bioline). the pcr amplification products were visualized using gel electrophoresis on 2% agarose, and staining with 0.5 µg ml–1 ethidium bromide for 10 min and documented with a gel doc xr camera using quantity one 4.6.1 software (bio-rad, hercules, ca, usa). 16s rdna amplification was performed using primers 27f and 809r (jungblut et al., 2005) with an initial denaturation step at 92 oc for 2 min followed by 35 cycles of 94°c for 10 s, 60°c for 20 s and 72°c for 1 min and a final extension step at 72°c for 5 min (jungblut et al., 2005). m. aeruginosa pcc7806 was used as positive control. hepatotoxin (hep) pcr reactions were performed using primers hepf and hepr targeting mcye/ndaf gene (jungblut and neilan, 2006). an initial denaturation step at 92°c for 2 min was followed by 35 cycles of 92°c for 20 s, 52°c for 30 s, and 72°c for 1 min, with a final extension step at 72°c for 5 min. the pcr products were sent to ramaciotti centre for genomics (university of new south wales, sydney australia) and sequencing was performed using the illumina miseq platform (illumina, san diego, ca, usa). using a pandaseq (ver. 2.4) nucleotide sequence were reconstructed (masella et al, 2012). overlapping regions were tab. 1. features of the studied lakes. waterbody surface area common use dominant cyanobacteria max. depth i̇znik lake 300 km2 recreation irrigation nodularia spumigena 65 m planktothrix rubescens cylindrospermopsis raciborskii dolichospermum sp. anabaenopsis sp. sapanca lake 46.8 km2 drinking water 55 m recreation planktothrix rubescens küçükçekmece lagoon 15.22 km2 recreation microcystis aeruginosa 20 m planktothrix agardhii microcystis wesenbergii manyas lake 159 km2 fisheries activities microcystis aeruginosa 3.4 m recreation microcystis wesenbergii irrigation sphaerospermopsis sp. dolichospermum flos-aquae cuspidothrix issatschenkoi taşkısı lake 0.75 km2 fisheries activities microcystis spp. 4.5 m dolichospermum sp. no nco mm er cia l u se on ly molecular identification of hepatotoxic cyanobacteria in turkey 55 aligned and scored. sequences were identified using the blastn search program (ncbi). hepatotoxin analysis microcystin/nodularin production of environmental blooms and isolated strains were measured by high performance liquid chromatography (hplc) with photodiode array (pda) detector (perkin elmer, usa) according to lawton (1994). lyophilized samples (10-50 mg) were extracted in 70% (v/v) aqueous methanol with ultrasonication and centrifuged at 14,000 x g for 5 min. clear supernatants were injected into the hplc column (waters symmetry c18, 3.9 × 150 mm, 5 μm particle size). elution mode was used: injection volume 25 µl, flow rate 1 ml min–1 and column temperature 40°c. mobile phases were milli-q water and acetonitrile both containing 0.1% (v/v) tfa. eluent absorbance was monitored from 200 to 300 nm and microcystins were detected at 238 nm. the limit of detection was 0.4 ng per injection corresponding to 0.001 µg mg−1 dw. results cyanobacteria species species identification was done by microscopy. since all blooms were mainly dominated by a single species,16s rdna results are very well correlated with microscopical examination. cyanobacteria that belong to three orders, chroococcales, nostocales and oscillatoriales, were detected. a total of nine species, anabaenopsis elenkinii v.v. miller, cylindrospermopsis raciborskii (woloszynska) seenayya & subba raju, sphaerospermopsis aphanizomenoides (previously denominated aphanizomenon aphanizomenoides forti), n. spumigena, m. aeruginosa, microcystis flos-aquae (wittrock) kirchner, microcystis wesenbergii (komárek) komárek ex komárek, p. rubescens, and planktothrix agardhii (gomont) anagnostidis & komárek were identified. the 16s rdna gene sequences obtained from both strains and environmental samples were assigned using blastn search of the natab. 2. hplc and hep pcr results for environmental bloom samples. code dominant species* place of collection date of collection hplc results hep pcr genbank accession (µg mg–1 d.w) results numbers e1 planktothrix agardhii küçükçekmece lagoon 27/10/2004 nd ky091680 e2 planktothrix agardhii küçükçekmece lagoon 03/11/2004 nd + ky091681 e3 planktothrix agardhii küçükçekmece lagoon 11/11/2004 nd ky091682 e4 microcystis aeruginosa küçükçekmece lagoon 04/10/2006 2.9 + ky091683 e5 planktothrix rubescens sapanca lake 06/02/2007 6.0 + ky091684 e6 planktothrix rubescens sapanca lake 21/02/2007 4.7 + ky091685 e7 microcystis aeruginosa küçükçekmece lagoon 28/09/2007 nd + ky091686 e8 planktothrix rubescens sapanca lake 23/01/2008 0.3 + ky091687 e9 anabaenopsis elenkinii i̇znik lake 16/05/2008 nd ky091688 e10 planktothrix rubescens sapanca lake 28/01/2009 1.1 + ky091689 *species: according to microscopic identification; nd, not detected. tab. 3. hplc and hep pcr results for cyanobacterial cultures. code cyanobacterial species* origin strain hplc results hep pcr genbank accession (µg mg–1 d.w) results numbers s1 microcystis aeruginosa küçükçekmece lagoon ifcc-ma03 6.8 + ky077257 s2 microcystis flos-aquae küçükçekmece lagoon ifcc-mf01 nd ky077258 s3 microcystis wesenbergii taşkısı lake ifcc-mw01 2.4 + ky077259 s4 anabaenopsis elenkinii i̇znik lake ifcc-ae01 nd ky077260 s5 sphaerospermopsis aphanizomenoides i̇znik lake ifcc-aa05 nd ky077261 s6 sphaerospermopsis aphanizomenoides i̇znik lake ifcc-aa01 nd ky077262 s7 cylindrospermopsis raciborskii manyas lake ifcc-cr01 nd ky077263 s8 nodularia spumigena i̇znik lake ifcc-ns01 3.2 + ky077264 s9 nodularia spumigena i̇znik lake ifcc-ns03 3.0 + ky077265 s10 planktothrix agardhii küçükçekmece lagoon ifcc-pa01 nd ky077266 s11 planktothrix rubescens sapanca lake ifcc-pr04 4.3 + ky077267 *species: according to microscopic identification; nd, not detected. no nco mm er cia l u se on ly l. köker et al.56 tional biotechnology information (ncbi) database (http://ncbi.nlm.nih.gov/blast/) (tabs. 2 and 3). the blast search showed 98-100% similarities. detection of hepatotoxin genes the hep pcr reaction resulted in amplification of a fragment in the expected size from two of three microcystis sp. strains, p. rubescens and two n. spumigena strains. no pcr product was obtained from strains assigned to p. agardhii, c. raciborskii, a. elenkinii and s. aphanizomenoides (tab. 3). the hep fragment was successfully amplified from five of seven planktothrix sp., one of two m. aeruginosa dominated environmental bloom samples. in culture samples, m. aeruginosa (s1) and m. flosaquae (s2) strains were isolated from same bloom recorded in küçükçekmece lagoon. while m. aeruginosa strain showed a hep-pcr product, m. flos-aquae was found negative (tab. 3). the other microcystis morphospecies, m. wesenbergii gave a positive result and showed hep-pcr product. the nostocalen species; s. aphanizomenoides and a. elenkinii did not give positive result as well as c. raciborskii strain. in environmental samples, the hep pcr reactions resulted in amplification of a 472-bp fragments for eight of ten samples. the mcye products were obtained from one of three p. agardhii bloom sample (e2), while no pcr products were obtained from p. agardhii (e1-e3) bloom samples. pcr-amplification of the amt domain was succesfully attained from all p. rubescens samples. to verify that the resulting amplicons, all pcr–amplified products from various lakes were sequenced. blast searches were used to identify similar sequences from genbank. detection of hepatotoxins cyanobacterial hepatotoxins were detected by hplcpda. total microcystin concentrations varied from 0.3 to 6.8 microcystin-lr equivalents µg mg–1 d.w. (tabs. 2 and 3). nodularin concentrations in ifcc-ns01 (s8) and ifcc-ns03 (s9) were 3.2 and 3.0 µg mg–1, respectively. the highest amount of microcystin (6.8 µg mg–1 d.w.) was found in m. aeruginosa (s1) strain. microcystin content of m. wesenbergii (s3) was found to be 2.4 µg mg–1. hplc analyses confirmed no microcystin presence in m. flos-aquae (s2), a. elenkinii (s4), s. aphanizomenoides (s5, s6), c. raciborskii (s7) and p. agardhii (s10) strains. in environmental samples, microcystins were not detected in p. agardhii (e1, e3) and a. elenkinii (e9) bloom samples. while mcye products were obtained from p. agardhii (e2) and m. aeruginosa (e7), microcystin was not detected by hplc. microcystin content of p. rubescens samples varied between 0.3-6 µg mg–1. discussion cyanobacteria species were shown to be the main component of phytoplankton community in lakes and reservoirs. earlier records on the algal flora of turkish waterbodies reported taxonomic lists, which were based on the microscopical monitoring and showed a diverse cyanobacteria community (aykulu and obalı, 1981; fakıoğlu et al., 2011). however, polyphasic approaches in classification of organisms are essential, since morphological characters are often unstable and incongruent with molecular tools. for example, the genus microcystis has several morphospecies sharing rather similar characteristics and discussions on the taxonomic assignment of these morphotypes is ongoing (bittencourt-oliveira, 2003). within the genus microcystis, typically two morphospecies (m. aeruginosa and m. flos-aquae) are found in the same population. according to the results of molecular methods used in this study, the mcye gene occurred in m. aeruginosa (s1) strain, but not in m. flos-aquae isolated from the same bloom. tillett et al. (2001) also did not find mcya gene occurrence among m. flos-aquae strains. however, mcya and b genes were detected in half of the colonies assigned to m. flos-aquae (total number was 8) isolated from lakes in europe. correspondingly, m. aeruginosa (n=149) had a higher proportion of colonies containing the mcya/b gene (via-ordorika et al., 2004). in this study, the third strain of microcystis isolated from taşkısı lake was assigned to m. wesenbergii (s3) and not only it contained the mcye gene but also produced microcystin (2.4 µg mg–1 d.w.). according to the study of viaordorika et al. (2004) this morphospecies was found non-toxic in all colonies (n=21) from european lakes. maršálek et al. (2001) showed that in czech republic m. wesenbergii contains little or no microcystin, similarly no microcystin was detected in colonies isolated from a czech reservoir (welker et al., 2007). also, molecular and chemical analysis did not show microcystin production in 250 individual colonies and 21 strains of m. wesenbergii isolated from chinese lakes (xu et al., 2008). however, otsuka et al (1999) found that m. wesenbergii has toxic and nontoxic strains. yosuno et al. (1998) also found that all m. wesenbergii (n=8) strains examined contained microcystin. likewise, in lake kastoria (greece), m. wesenbergii dominant bloom containing toxin producing genes such as mcya and mcyb was reported (gkelis et al., 2014). pavlova et al. (2014; 2015) found toxic bloom dominated by m. wesenbergii in lake dourankoulak, and highlighted that toxicity may vary between clones of the same strain. because of these contradictory results, it is necessary to analyse higher number of microcystis morno nco mm er cia l u se on ly molecular identification of hepatotoxic cyanobacteria in turkey 57 phospecies to determine the relationship between toxigenicity and morphological characters. it is known that p. agardhii and p. rubescens have specific ecological niches. while p. rubescens occurs in oligoto mesotrophic physically stratified lakes (akçaalan et al., 2014a), p. agardhii become dominant in shallow, eutrophic and polymictic water bodies (kurmayer et al., 2004). in this study, p. rubescens was isolated from sapanca lake, which is a moderately deep, oligo-mesotrophic lake. in contrast, p. agardhii formed a bloom in a hypereutrophic lake in late autumn and polymictic conditions. similar to microcystis both toxic and nontoxic strains can be found in the same population of p. agardhii and p. rubescens (kurmayer et al., 2004; akçaalan et al., 2006). in general, the share of strains containing the mcya/b gene is highest in p. rubescens populations in contrast to p. agardhii. accordingly, our results showed that p. rubescens has active microcystin genes, while the strain isolated from p. agardhii bloom was found nontoxic. the strain of a. elenkinii was isolated from a bloom sample of i̇znik lake which was dominated by this species. both the bloom sample and isolated strain were found negative for the mcy genes as well as no microcystin was detected by hplc. this species generally cooccurs with other nostocalen cyanobacteria and toxicity is attained to all of them (maršálek et al., 2000; papadimitriou et al., 2013). however, there is no record of microcystin production of a isolated strain of a. elenkinii. c. raciborskii has been shown to produce hepatotoxic cylindrospermopsin and neurotoxic saxitoxins (wood and stirling, 2003; molica et al., 2005). this species originates from tropical regions and currently expands its distribution in temperate regions, therefore it may be considered an invasive species in european waterbodies (padisák, 1997; moreira et al., 2015). in this study c. raciborskii was isolated from shallow hypereutrophic manyas lake but did not contain the mcye gene. also, no cylindrospermopsin was detected according to molecular and analytical analysis (data not shown). there are some contradictory results between molecular and analytical methods. m. aeruginosa (e5) and p. agardhii (e2) contained the mcye gene, but did not produce microcystin as revealed by hplc. studies showed that cyanobacteria strains with mcy genes lacked detectable microcystins as a result of inactivation of the genes (neilan et al., 1999; nishizawa et al.,1999; kaebernick et al., 2001; tillett et al., 2001; mikalsen et al., 2003). samples used in this study were collected from waterbodies with different morphological and physicochemical characteristics. some cyanobacteria species have been found in both shallow and moderately deep lakes, some others prefer deep waterbodies. however, the distribution of species is governed mainly by trophic situation of the lakes. microcystis species together with p. agardhii formed blooms in eutrophic environment, such as manyas, küçükçekmece and taşkısı lake. nostocalen cyanobacteria species, on the other hand, prefer alkaline, meso-eutrophic waters of i̇znik lake (akçaalan et al., 2009, 2014b). especially nodularia spumigena is an euryhaline species living in hyposaline to brackish waters in turkey (kocasari et al., 2015; kızılkaya et al., 2016). similarly, a. elenkinii is also known as a hyposaline species (kemp, 2009; kotut and krienitz, 2011). the growth of these species might have been supported by high conductivity of the lake water. on the other hand, in typical freshwater sapanca lake, which is used for drinking water and has low nutrient concentration, toxic p. rubescens form massive blooms. the most important factors are the high water transparency, thermal stratification, a long water residence time and low nutrient availability, which have negative effect on other phytoplankton species in the lake (legnani et al., 2005; akçaalan et al., 2014a) conclusions in conclusion, applications of molecular and dna amplification methods provide a great advantage for monitoring toxic cyanobacterial blooms in the aquatic environments. it has a potential to identify the organisms and to detect their cyanotoxin production. this study, using different methods collaboratively, shows that toxic cyanobacteria blooms are very common in turkish inland waterbodies with different trophic levels. to our knowledge, this is the first detailed study identifying different toxic cyanobacteria species and their hepatotoxin production in turkey using molecular methods. acknowledgments this work was supported by scientific research projects coordination unit of istanbul university. project number 2846. we would like to thank ban group at the university of new south wales for their help in molecular work. the authors also would like to acknowledge the european cooperation in science and technology, cost action es 1105 “cyanocost” for adding value to this study through networking and knowledge sharing with european researchers. references akçaalan r, young fm, metcalf js, morrison lf, albay m, codd ga, 2006. microcystin analysis in single filaments of planktothrix spp. in laboratory cultures and environmental blooms. water. res. 40:1583-1590. no nco mm er cia l u se on ly l. köker et al.58 akçaalan r, marzur-marzec h, zalewska a, albay m, 2009. phenotypic and toxicological characterization of toxic nodularia spumigena from a freshwater lake in turkey. harmful algae 8:273-278. akçaalan r, köker l, gürevin c, albay m, 2014a. planktothrix rubescens: a perennial presence and toxicity in lake sapanca. turk. j. bot. 38:782-789. akçaalan r, köker l, oğuz a, spoof l, meriluoto j, albay m, 2014b. first report of cylindrospermopsin production by two cyanobacteria (dolichospermum mendotae and chrysosporum ovalisporum) in lake iznik, turkey. toxins 6:31733186. akçaalan r, köker l, albay m, 2016. do eutrophic waters prompt to toxic cyanobacteria in turkish black sea coast? j. environ. prot. ecol. 17: 584-592. albay m, akçaalan r, tufekci h, metcalf js, beattie ka, codd ga, 2003a. depth profiles of cyanobacterial hepatotoxins (microcystins) in three turkish freshwater lakes. hydrobiologia 505:89-95. albay m, akçaalan r, aykulu g, tufekci h beattie ka, codd ga, 2003b. occurrence of toxic cyanobacteria before and after copper sulphate treatment in a water reservoir, istanbul, turkey. algol. stud. 109:67-78. albay m, matthiensen a, codd ga, 2005. occurrence of toxic blue-green algae in the küçükçekmece lagoon (istanbul, turkey). environ. toxicol. 20:227-284. aykulu g, obalı o, 1981. phytoplankton biomass in the kurtboğazı dam lake. com. de fac. sci. univ. ankara c2 24:30-45. anagnostidis k, komárek j, 1988. modern approach to the classification system of cyanophytes. algol. stud./arch. hydrobiol. 50-53:327-472. aykulu g, dogan k, hasirci s, 1999. a study on the phytoplankton communities of lakes taskisi and poyrazlar (adapazari, turkey). istanbul university j. aquat. prod. special issue 157-184. azevedo, smfo, carmichael ww, jochimsen em, rinehart kl, lau s, shaw gr, eaglesham gk, 2002. human intoxication by microcystins during renal dialysis treatment in caruaru-brazil. toxicology 181-182:441-446. backer lc, manassaram-baptiste d, leprell r, bolton b, 2015. cyanobacteria and algae blooms: review of health and environmental data from the harmful algal bloom-related illness surveillance system (habiss) 2007-2011. toxins 7:1048-1064. bittencourt-oliveira m, 2003. detection of potential microcystin-producing cyanobacteria in brazilian reservoirs with a mcyb molecular marker. harmful algae 2:51-60. bláha l, babica p, maršálek b, 2009. toxins produced in cyanobacterial water blooms-toxicity and risks. interdiscip. toxicol. 2:36-41. bortoli s, volmer da, 2014. account: characterization and identification of microcystins by mass spectrometry. eur. j. mass spectrom. 20:1-19. bukowska a, bielczynska a, karnkowska a, chrost rj, jasser i, 2014. molecular (pcr-dgge) versus morphological approach: analysis of taxonomic composition of potentially toxic cyanobacteria in freshwater lakes. aquat biosyst. 10:2. carmichael ww, azevedo smfo, an js, molica rjr, jochimsen em, lau s, rinehart kl, shaw gr, eaglesham gk, 2001. human fatalities from cyanobacteria: chemical and biological evidence for cyanotoxins. environ. health. persp. 109:663-668. çelik k, ongun t, 2006. seasonal dynamics of phytoplankton assemblages across nutrient gradients in the shallow hypertrophic lake manyas, turkey. lake reserv. manage. 22:250-260. codd ga, bell sg, kaya k, ward cj, beattie ka, metcalf js. 1999. cyanobacterial toxins exposure routes and human health. eur. j. phycol. 34:405-415. fakıoğlı ö, atamanalp m, demir n, 2011. toxic blue-green algae in dam lakes. ankara üniversitesi çevrebilimleri 3:65-71. gkelis s, zaoutsos n, 2014. cyanotoxin occurrence and potentially toxin producing cyanobacteria in freshwaters of greece: a multi-disciplinary approach. toxicon 78:1-9. harke mj, steffen mm, gobler cj, otten tg, wilhelm sw, wood sa, paerl hw, 2016. a review of the global ecology, genomics, and biogeography of the toxic cyanobacterium, microcystis spp. harmful algae 54:4-20. jungblut ad, hawes i, mountfort d, hitzfeld b, dietrich dr, burns bp, neilan ba, 2005. diversity within cyanobacterial mat communities in variable salinity meltwater ponds of mcmurdo ice shelf, antarctica. environ. microbiol. 7:519-529. jungblut ad, neilan ba, 2006. molecular identification and evolution of the cyclic peptide hepatotoxins, microcystin and nodularin, synthetase genes in three orders of cyanobacteria. arch. microbiol. 185: 107-114. kaebernick m, rohrlack t, christoffersen k, neilan ba., 2001. aspontaneous mutant of microcystin biosynthesis: genetic charac-terization and effect on daphnia. environ. microbiol. 3:669-679. kemp as, 2009. freshwater cyanoprokaryota blooms in the swan coastal plain wetlands: ecology, taxonomy and toxicology. phd thesis, curtin university of technology, australia. kızılkaya it, demirel z, kesici k, kesici e, sukatar a, 2016. morphological, molecular and toxicologial characterization of nodularia spumigena mertens in jungens (1822) from brackishwater lake bafa (turkey). sinop uni j nat sci 1:39-52. kocasari fs, gulle i, kocasari s, pekkaya s, mor f, 2015. the occurrence and levels of cyanotoxin nodularin from nodularia spumigena in the alkaline and salty lake burdur, turkey. j. limnol. 74:530-536. komárek j, 2013. cyanoprokaryota. 3. heterocytous genera. springer verlag, berlin: 1130 pp. komárek j, anagnostidis k, 1986. modern approach to the classification system of cyanophytes. algol. stud./arch. hydrobiol 43:157-164. komárek j, anagnostidis k, 1999. [cyanoprokaryota. 1. teil: chroococcales, p. 1-548]. in: h. ettl, g. gardner, h. heynig and d. mollenheuer (eds.),[süsswasserflora von mitteleurope].[book in german]. gustav fischer verlag. komárek j, anagnostidis k, 2005. [süßwasserflora von mitteleuropa, bd. 19/2: cyanoprokaryota].[book in german]. springer spektrum, heidelberg: 759 pp. kotut k., krienitz l, 2011. does the potentially toxic cyanobacterium microcystis exist in the soda lakes of east africa? hydrobiologia 664:219-225. no nco mm er cia l u se on ly molecular identification of hepatotoxic cyanobacteria in turkey 59 kurmayer r, christiansen g, fastner j, börner t, 2004. abundance of active and inactive microcystin genotypes in populations of the toxic cyanobacterium planktothrix spp. environ. microbiol. 6:831-841. kurmayer r, christiansen g, 2009. the genetic basis of toxin production in cyanobacteria. freshwater rev. 2:31-50. lawton la, edwards c, codd ga, 1994. extraction and highperformance liquid chromatographic method for the determination of microcystin in raw and treated waters. analyst 119:1525-1530. legnani e, copetti d, oggioni a, tartari g, palumbo mt, morabito g, 2005. planktothrix rubescens’ seasonal dynamics and vertical distribution in lake pusiano (north italy). j. limnol. 64:61-73. masella ap, bartram ak, truszkowski jm, brown dg, neufeld jd, 2012. pandaseq: paired-end assembler for illumina sequences. bmc bioinformatics 13:31. maršálek b, bláha l, hindák f, 2000. review of toxicity of cyanobacteria in slovakia. biologia 55:645-652. maršálek b, bláha l, turánek j, neča j, 2001. microcystin-lr and total microcystin in cyanobacterial blooms in the czech republic 1993-1998, p. 56-62. in: i. chorus (ed.), cyanotoxins-occurence, causes, consequences. springer-verlag. merel s, villarín mc, chung k, snyder s, 2013. spatial and thematic distribution of research on cyanotoxins. toxicon 76:118-131. bernard c, ballot a, thomazeau s, maloufi s, furey a, mankiewicz�boczek j, pawlik�skowrońska b, capelli c, salmaso n, 2017. cyanobacteria associated with the production of cyanotoxins, appendix 2. in: j. meriluoto, l. spoof, and g. codd (eds.), handbook of cyanobacterial monitoring and cyanotoxin analysis. j. wiley & sons. mikalsen b, boison g, skulberg om, fastner j, davies w, gabrielsen tm, rudi k, jakobsen ks., 2003. natural variationin the microcystin synthetase operon mcyabc and impact on microcystin production in microcystis strains. j. bacteriol. 185:2774-2785. moffitt mc, neilan, ba, 2001. on the presence of peptide synthetaseand polyketide synthase genes in the cyanobacterial genus nodularia. fems microbiol. lett. 196:207-214. molica rjr, oliveira eja, carvalho pvvc, costa ansf, cunha mcc, melo gl, azevedo smfo, 2005. occurrence of saxitoxins and an anatoxin-a(s)-like anticholinesterase in brazilian drinking water supply. harmful algae 4:743-753. moreira c, fathalli a, vasconcelos v, antunes a, 2015. phylogeny and biogeography of the invasive cyanobacterium cylindrospermopsis raciborskii. arch. microbiol. 197:47-52. neilan ba, dittmann e, rouhiainen l, bass ra, schaub v, sivonen k, borner t, 1999. nonribosomal peptide synthesisand toxigenicity of cyanobacteria. j bacteriol 181:4089-4097. neilan ba, jacobs d, deldot t, blackall ll, hawkins pr, cox pt, goodman ae, 1997. rrna sequences and evolutionary relationships among toxic and nontoxic cyanobacteria of the genus microcystis. internat. j. syst. bacteriol. 47:693-697. nishizawa t, asayama m, fujii k, harada k, shirai m., 1999. genetic analysis of the peptide synthetase genes for a cyclic heptapeptide microcystin in microcystis spp. j biochem (tokyo) 126:520-529. otsuka s, suda s, li r, watanabe m, oyaizu h, matsumoto s, watanabe mm, 1999. phylogenetic relationships between toxic and non-toxic strains of the genus microcystis based on 16s to 23s internal transcribed spacer sequence. fems microbiol. lett. 172:15-21. padisák j, 1997. cylindrospermopsis raciborskii (woloszynska) seenayya et subba raju, an expanding, highly adaptive cyanobacterium: worldwide distribution and review of its ecology. arch. hydrobiol. 107:563-593. paerl h, otten t, 2013. harmful cyanobacterial blooms: causes, consequences, and controls. microb. ecol. 65:995-1010. papadimitriou t, katsiapi m, kormas ka, moustaka-gouni mik, 2013. artificially-born “killer” lake: phytoplankton based water quality and microcystin affected fish in a reconstructed lake. sci. total. environ. 452-453:116-124. pavlova v, stoyneva m, georgieva v, donchev d, spoof l, meriluoto j, bratanova z, karadjova i, 2014. new records of microcystins in some bulgarian water bodies of health and conservational importance. j. water resource prot. 6:446-453. pavlova v, stoyneva-gärtner m, uzunov b, bratanova z, lazarova a, karadjova i, 2015. microcystins-lr, -yr andrr in six bulgarian water bodies of health and conservational importance (2012-2014). j. water resource prot. 7:1375-1386. rippka r, deruelles j, waterbury j, herdman m, stanier r, 1979. generic assignments, strain histories and properties of pure cultures of cyanobacteria. j. gen. microbiol. 111:1-61. sivonen k, niemelä si, niemi rm, lepistö l, luoma th, räsänen la, 1990. toxic cyanobacteria (blue green algae) in finnish fresh and coastal waters. hydrobiologia 190:267-275. sivonen k, jones g, 1999. cyanobacterial toxins, p. 41-111. in: i. chorus and j. bartam (eds.), toxic cyanobacteria in water: a guide to their public health consequences, monitoring and management. e & fn spon. spoof l, karlsson k, meriluoto j, 2001. high-performance liquid chromatographic separation of microcystins and nodularin, cyanobacterial peptide toxins, on c18 and amide c16 sorbents. j. chromatogr. a 909:225-236. tas b, gonulol a, 2007. an ecologic and taxonomic study on phytoplankton of a shallow lake, turkey. j. environ. biol. 28:439-445. tillett d, neilan ba, 2000. xanthogenate nucleic acid isolation from cultured and environmental cyanobacteria. j. phycol. 36:251-258. tillett d, parker dl, neilan ba, 2001. detection of toxigenicity by a probe for the microcystin synthetase a gene (mcya) of the cyanobacterial genus microcystis: comparison of toxicities with 16s rrna and phycocyanin operon (phycocyanin intergenic spacer) phylogenies. appl. environ. microb. 67:2810-2818. via-ordorika l, fastner j, kurmayer r, hisbergues m, dittmann e, komárek j, erhard m, chorus i, 2004. distribution of microcystin-producing and non-microcystin-producing microcystis sp. in european freshwater bodies: detection of microcystins and microcystin genes in individual colonies. syst. appl. microbiol. 27:592-602. welker m, šejnohová l, némethová d, von dohren h, jarkovsky j, maršálek b, 2007. seasonal shifts in chemotype composition of microcystis sp. communities in the pelagial and the sediment of a shallow reservoir. limnol. oceanogr. 52:609-619. westrick ja, szlag dc, southwell bj, sinclair j, 2010. a review no nco mm er cia l u se on ly l. köker et al.60 of cyanobacteria and cyanotoxins removal/inactivation in drinking water treatment. anal. bioanal. chem. 397:17051714. whitton ba, potts m, 2007. the ecology of cyanobacteria: their diversity in time and space. springer science, berlin: 669 pp. wood sa, stirling dj, 2003. first identification of the cylindrospermopsin-producing cyanobacterium cylindrospermopsis raciborskii in new zealand. new zeal. j. mar. fresh. 37:821-828. xu y, wu z, yu b, peng x, yu g, wei z, wang g, li r, 2008. non-microcystin producing microcystis wesenbergii (komárek) komárek (cyanobacteria) representing a main waterbloom-forming species in chinese waters. environ. pollut. 156:162-167. yosuno m, sugaya y, kaya k, watanabe mm, 1998. variations in the toxicity of microcystis species to moina macrocopa. phycol. res. 46:31-36. no nco mm er cia l u se on ly layout 1 introduction oxbow lakes are important components of the floodplain systems of lowland rivers. rivers supply these lakes with suspended matter and nutrients during flood events, while during inter-flood periods even oxbow lakes located close to the river become isolated and develop the characteristics of lentic ecosystems. as a result, oxbow lakes behave alternatively as lotic or lentic ecosystems in relation to the hydrological regime of the river. flood events in oxbow lakes can be reconstructed by sediment analyses by distinguishing lotic and lentic regimes (wolfe et al., 2006). zooplankton biodiversity and abundance is commonly poor in lotic environments (during floods), in comparison to lentic waters (between floods) (zsuga, 1998, 1999). these conditions are also well reflected by the remains of zooplankton species, especially cladocerans, which accumulate in sediment deposits. as a consequence, subfossil cladocera serve as a useful biological proxy for the evaluation of lake responses to environmental changes such as water level changes and trophic structure (korhola and rautio, 2001; korponai et al., 2010a, 2011; jeppesen et al., 2011). in hungary, a study by korponai et al. (2011) on cladoceran remains was used for reconstructing past changes in trophic status of lake balaton. goslar et al. (1999) found high frequency of bosmina longirostris (o.f. müller, 1785) during the german colonization of lake gościąź, and a decrease of this species after the demolition of the settlement. as b. longirostris prefers eutrophic conditions (korponai et al., 2010a, 2011), the authors concluded that the german settlers induced an increase in the lake’s eutrophic status, which is in line with results provided by other sediment proxies. galbarczyk-gąsiorowska et al. (2009) could relate changes in cladoceran communities to decreasing water level in stare biele mire in poland. planktonic cladocerans gradually disappeared while advances in oceanography and limnology, 2016; 7(2): 131-141 article doi: 10.4081/aiol.2016.6168 this work is licensed under a creative commons attribution-noncommercial 4.0 international license (cc by-nc 4.0). reconstruction of flood events in an oxbow lake (marótzugi-holt-tisza, ne hungary) by using subfossil cladoceran remains and sediments jános korponai,1,2* istván gyulai,3 mihály braun,4 csilla kövér,1 istván papp,5 lászló forró6 1mta pe limnoecology research group, egyetem u. 10, 8200 veszprém; 2department of chemistry and environmental sciences, university of west hungary, károly gáspár tér 4, 9700 szombathely; 3institute of biology and ecology, university of debrecen, egyetem tér 1, 4032 debrecen; 4laboratory of environmental studies, institute for nuclear research, hungarian academy of sciences, bem tér 18/c, 4026 debrecen; 5department of mineralogy and geology, university of debrecen, egyetem tér 1., 4032 debrecen; 6department of zoology, hungarian natural history museum, baross u. 13, 1088 budapest, hungary *corresponding author: korponai.janos@iif.hu abstract oxbow lakes are important components of the floodplain systems of lowland rivers. during flood events, oxbows are connected with the main river channel, and behave as lotic systems, while during inter-flood periods, these lakes can be considered as lentic ecosystems. rivers are generally poor in planktonic organisms and their sediments contain scarce biological remains in comparison to lentic water ecosystems. however, due to their alternating running and standing water regime, sedimentary biological remains of oxbow lakes can be used as proxies for tracking changes of past hydrological regimes. in this study we investigated how cladoceran communities respond to flood events, and whether flood events can be recognized by community analysis of cladoceran remains. a sediment core from marótzugi-holt-tisza oxbow lake was analyzed for identification of past flood events based on changes in the subfossil cladocera community. floods were defined based on the proportion of fine sand (50 µm grain size) in the oxbow sediments. if the fine sand portion was <3%, the water regime of the oxbow was considered as lentic, otherwise it was lotic. both organic and pigment contents were significantly higher in the core sections deposited during lentic stages. thirty-four cladocera species were determined in this core, all common to littoral habitats of eutrophic shallow lakes in hungary. one planktonic (bosmina longirostris) and four chydorid species (alona rectangula, acroperus harpae, alonella nana and chydorus sphaericus) were dominant throughout the core and contributed >90% of total remains. discriminant analysis on cladoceran data confirmed that lotic and lentic hydrological stages were characterized by different cladocera species associations. bosmina longirostris, chydorus sphaericus, alona rectangula, acroperus harpae, leydigia leydigi, a. quadrangularis and a. nana were mainly responsible for the differences between lotic and lentic species assemblages. our results revealed that cladocera remains can be used to track changes in the hydrological regime of oxbow lakes. key words: multiproxy reconstruction; oxbow lakes; lotic/lentic regimes; loi; sedimental pigments; subfossil cladocera. received: july 2016. accepted: november 2016. no n c om me rci al us e o nly 132 j. korponai et al. macrophyte-associated species appeared and increased until finally a poor cladoceran community developed, which consisted of very rare species able to tolerate specific conditions (i.e., low ph, low nutrient levels) in the peat phase. according to luoto et al. (2011), lotic conditions strongly affect cladoceran communities of boreal shallow lakes, so that sediment layers deposited during lotic stages may introduce errors in temperature and water-depth reconstructions. in oxbow lakes trophic status is typically higher between flood events (knowlton and johns, 1997). however, as cladoceran distribution can be affected by different abiotic factors, we hypothesize that flood events could also play an important role in driving changes in cladoceran communities of oxbows. the aim of this study is to verify the occurrence of past flood events through the analysis of cladoceran remains, and to reconstruct the effects of flood events on cladoceran communities. methods study area marótzugi-holt-tisza (holt=dead) is a small unprotected oxbow lake (length 1.8 km, width 60 m, area 10 ha) located at the left side of the river tisza (at the 568th river km), close to the gávavencsellő village in ne hungary (n 48.175611, e 21.612306, fig. 1). the river tisza is one of the largest rivers in central europe, and the largest water flow of the hungarian plain. it is 946 km long, and the catchment area is 157,186 km2. the water regime of the river tisza is very variable. the highest water levels occur at the time of snow melt between february and april. a second high water level stage frequently occurs in june, due to intense seasonal rainfalls. the lowest water levels are commonly measured from august to the end of september. from an economical point of view, the tisza is the most important river in hungary, since it provides the largest irrigation water supply for cultivated areas in the country (234-265 kha in 1970-80, lászlóffy, 1982). before its regulation, the river tisza was surrounded by large floodplains covering a total area of 19,637 km2, 4770 km2 of which were permanently inundated. therefore, the tisza floodplain occupied 21% of the country, while 5% was permanently inundated (botár and károlyi, 1971). in the late 19th century, a complex drainage system was constructed, consisting of floodgates, artificial river beds and channels. meanders were cut off, reducing the length of the river tisza from 1420 km to 946 km. altogether 112 meanders were isolated from the river-bed and fig. 1. map showing the location of the marótzugi-holt-tisza and the coring point. no n c om me rci al us e o nly reconstruction of flood events in an oxbow lake 133 converted into oxbow lakes (lászlóffy, 1982). by cutting the bends, the slope of the river-bed increased, thus increasing current velocity and kinetic energy of the river. this was supposed to prevent sediment build-up in the riverbed, as the rates of erosional and depositional processes became balanced. according to the present nomenclature, oxbows lying within the river floodplain are called unprotected (plesiopotamon), while those that are located outside of levees are defined as protected (paleopotamon). unprotected oxbows are frequently affected by floods therefore their sediments become laminated due to sedimentation of organic matter following floods. river canalization created 80 oxbow lakes from the river tisza. most of them are unprotected and highly affected by floods, which makes them ideal water ecosystems for studies on of flood events. previous studies on the sediments of the river tisza oxbows revealed that heavy metal pollution was a regular occurrence due to ore mining in maramures county in romania (braun et al., 2000, 2010). studies by korponai et al. (2010b, 2010c) showed that cladocera communities represent an outstanding portion of biodiversity of oxbows lakes. following the reclamation scheme, a large bend of the river tisza was isolated in 1860 and the marótzugi-holttisza oxbow was established (pálfai, 2001). this oxbow lake was chosen as a study site due to its known age and its favorable logistics. marótzugi-holt-tisza oxbow is unprotected and situated between dikes and the river channel. the biota of this lake is particularly rich, and has been registered as a national nature reserve and wildlife sanctuary (hortobágy national park). moreover, it was selected as part of the pilot project area for the phare-sponsored hungarian national biodiversity monitoring program (müller et al., 2000). coring a 463 cm long undisturbed sediment core was collected using a rod-operated piston corer from the deepest part of the oxbow (~2m depth, fig. 1) in spring in 2009. subsamples (1 cm3) were taken at 2 cm intervals and analyzed for grain size distribution, organic content, subfossil photosynthetic pigments and cladocera remains. sediment chronology historical records of major flood events at river tisza were used to determine the chronology of the sediment core. a public database of stream flow of the river tisza is available at http://www.hydroinfo.hu. daily stream flow has been recorded since 1856. the largest floods were determined by selecting events exceeding the highest flood warning level (iii level: 800 cm) at the staff gage at vásárosnamény. based on the time of establishment, the age of the core bottom was set at 1860. the next time horizon is the trace of a 1888 flood, which was one of the highest ever-recorded floods of river tisza (nyárády, 1900; lászlóffy, 1982; braun et al., 2000). braun et al. (2010) drilled a long sediment core in a nearby site of marótzugi-holt-tisza in 1997, and published a 137cs based chronology for marótzugi-holt-tisza. we could successfully apply the same 137cs chronology to our sediment core after parallelizing the loi profiles of the two cores. grain size and subfossil pigment analyses to determine the grain size distribution, sediment subsamples were oven-dried at 105°c for at least 24 h before recording dry mass. dried and grinded samples were mixed with 0.5 ml 0.002 m sodium-oxalate, then 2 ml concentrated glycerin were added. the samples were analyzed for particle size using a malvern mastersizer 2000. this system measures volumetric grain sizes between 0.12000 µm by analyzing in situ the angle of refraction of a laser beam that is aimed through the container holding the sediment-diluent mix. flood events were arbitrary defined as sediment layers containing more than 3% of fine sand (50 µm grain size). when the fine sand portion was less than 3% we defined the water regime of the oxbow as lentic, otherwise it was lotic. organic sediment content was determined as lost on ignition (loi) according to standardized methods (heiri et al., 2001). subfossil photosynthetic pigments were extracted with acetone and determined spectrophotometrically as chlorophyll degradation residues (spdu, subfossil pigment degradation units). the absorbance of the liquid phase was measured at 666 and 750 nm after sedimentation. the weight of the solid phase was measured after evaporating the acetone and drying until constant weight. spdu values were calculated according to vallentyne (1955). subfossil cladoceran subsamples for the subfossil cladoceran analysis were deflocculated in 10% koh. after the koh treatment, sediment samples were treated with 10% hcl in order to remove carbonaceous particles, and with 40% hf (for 2 h) to eliminate sand and clay fractions (frey, 1986). cladocera remains were collected by sieving through a 35 μm mesh (frey, 1986). only well preserved chitinous remains (headshields, carapaces, post-abdomens, post-abdominal claws, and ephippia) were considered to determine the density of cladocera species. fragments were counted only if unambiguous diagnostic marks were evident. the most frequent body parts of each taxon were used to estimate of the abundance of individuals as density (n. ind cm–3 of fresh sediment). the composition of the cladocera community was estimated based on the determination of at no n c om me rci al us e o nly 134 j. korponai et al. least 300 individuals in each subsample (korhola and rautio, 2001). taxonomical identification was carried out according to frey (1950, 1962, 1988, 1991), goulden and frey (1963), gulyás and forró (1999), sebestyén (1965, 1969, 1970, 1971), szeroczyńska and sarmaja-korjonen (2007), and whiteside et al. (1978). statistical analyses statistical analyses were carried out on cladocera logtransformed (log(x+1)) density data. we used hierarchical clustering on transformed data to reveal temporal changes in species structure. after calculating the euclidean distances, the ward agglomerative cluster analysis was applied. constrained cluster analysis was done to obtain homogeneous cladocera assemblage zones. clusters were formed on the basis of minimal within-group squared euclidean distances between objects (coniss, grimm, 1987). significant stratigraphic zones were determined by applying the broken stick model on coniss clustering. the classification of stages characterized by different water regime (lotic and lentic) was tested by linear discriminant analysis (lda) applied to transformed cladoceran data. differences in loi, spdu and lda during lotic and lentic stages were tested by kruskall-wallis (k-w) test. similarity percentage (simper) analysis based on braycurtis dissimilarity index was applied to outline cladoceran species responsible for community changes during lotic and lentic water regimes. lowess smoothing was applied for discriminant scores with 0.1 span. all statistical analyses were performed in the r statistical programming language (r development core team, 2010) using the rioja (juggins, 2009) and vegan (oksanen et al., 2016) packages. results sediment description and lithology the entire length of the sediment sequence was 463 cm, and four sections were distinguished on the basis of their different aspect (fig. 2). the oldest samples (463-201 cm depth) were characterized by alternating layers of coarse fig. 2. stratigraphy plot for medium and fine sand fractions in grain size distribution, loss on ignition (loi), sediment pigment degradation units (spdu), total cladocera abundances and discriminant (lda) scores in the sediment core of marótzugi-holt-tisza oxbow. red line: lowess smoothing of ld scores with 0.1 span; bold numbers indicate time horizons: 1888 largest historical flood; 1968 and 1986 according to 137cs based dating from braun et al. (2010), numbers in italics refer to inferred dates of major floods. no n c om me rci al us e o nly reconstruction of flood events in an oxbow lake 135 sand and clay silt. organic matter contents were low (around 2%) in the deepest layers (463-431 cm), which contained coarse calcareous concretions between 459-454 cm (fig. 2). from 431 cm to the top high clay content was found, generally accompanied by sandy and coarse layers of different thickness, and by thin black organic bands. organic matter (loi%) content varied between 3 and 9% of dry matter. two further subsections could be distinguished by loi (fig. 2). the lower subsection (431-130 cm depth) had lower loi content (~4%), while values increased up to 4-9% (mean=6%) above 130 cm depth. between 431 and 201 cm depth only coarse silt and sand layers were found (fig. 2). thick coarse sand and clay layers were found at 431-413, 410-380, and 375-369 cm depth, while thinner sand bands were recognized at 346348, and 331-330 cm. finally, an organic rich sand belt was found between 325 and 319 cm (fig. 2). thin sand bands were also recognized at 265, 255, 232, 210 cm (fig. 2). from 200 cm to the core top, eight thin black bands with high organic content were found at 127, 126, 125, 121, 116, 106, 101 and 100 cm, and further light brownish layers at 93-85, 73-70, and 68 cm. spdu content increased from 194 cm depth to the core top (fig. 2). average spdu was 0.167 1 g dm–1 between 462-190 cm and 3.439 g dm–1 from 190 cm to the core top. core chronology eleven large flood events were recognized in the hydrological history of the river tisza. the mark of a large disturbance in the sediment record was found at 454 cm as a calcareous concretion (fig. 2). digging new riverbed and isolating the river bend caused a large disturbance, and the corresponding sediment layer was therefore assumed as the time horizon of oxbow establishment in 1860. three thick coarse sandysilt layers at 431-413, 410-380, and 375-329 cm depth were interpreted as deposited between the 1870s and 1895. in fact, particularly large floods were historically recorded in 1876, 1881, 1888 and 1895. as peaks of medium size sand particle fractions mark the beginning of flood events (schweitzer et al., 2002), these layers were selected to indicate 1876, 1881, 1888 and 1895 years, respectively (fig. 2). sandy clay layers and peaks of medium size sand particle fractions at 340, 324, 315, 274, 258, 110, 70 and 42 cm were interpreted as markers of large historical floods in 1915, 1919, 1925, 1932, 1940, 1964, 1970 and 1979. the increase in loi and spdu values from 185 cm depth to the sediment top was interpreted as time horizon for the beginning of the building of a large dam and a reservoir close to the village tiszalök. this reservoir could stabilize the water regime of marótzugi oxbow, since the dam at tiszalök has a backwater effect in increasing the minimal water level which reaches the 628th river km. average sedimentation rates were estimated by applying a linear model to the core depth-flood profile. values were high, ranging between 3.05-3.62 cm y–1, below 75 cm depth, and gradually decreased down to 1.1 cm y–1 in the upper core section. during periods of enhanced flood frequency (i.e.,1881-95, 1932-40, 1940-50,1963-68, 1968-1970) sedimentation rates were particularly high, i.e., 4.3, 5.7, 7.3, 5.7 and 7 cm y–1 respectively. subfossil cladocera no cladocera remains were found in the deepest core section between 462 and 430 cm. very few remains were found in the sediment layers between 428-184 cm, while cladocera abundances gradually increased from the core bottom to the surface (fig. 2). overall, 34 cladocera species were found in the studied sediments sequence (fig. 3). all the species are characteristic for littoral habitats of eutrophic shallow lakes, with dense macrophytes belts. the hierarchical clustering grouped the identified cladoceran species according to their abundance into four clusters. only one species, the planktonic b. longirostris, was included in the first cluster, though remains of this species dominated the cladocera assemblages throughout the whole core (fig. 3). alona rectangula (sars, 1862) and chydorus sphaericus (o.f. müller, 1776) were the characterizing species of the second cluster. density of these euryecious species, which are very common in the majority of freshwaters, was high in layers between 110 and 70 cm depth. these species are. the third cluster included fourteen species, which were present throughout the core though with low abundances. rare species were grouped into the fourth cluster. four statistically significant zones were distinguished along the core on the basis of abundance of cladocera remains. cl-1 zone (462-184 cm) albeit a number of the deepest core layers were empty (fig. 2), remains of 30 cladoceran species were found in this zone (fig. 3). species abundances were very low (1-100 ind cm –3) except for b. longirostris, which reached a maximum of 300 ind cm–3 (fig. 2) at 312 cm. the zone was characterized by a variety of phytophilous chydorids, which were included in the second and third species cluster. pelagic b. longirostris was the dominant species, but c. sphaericus, a. rectangula, diparalona rostrata (koch, 1841), eurycercus lamellatus (o. f. müller, 1785), and acroperus harpae (baird, 1834) were also very abundant in this zone (fig. 3). cl-2 zone (184-112 cm) this zone was characterized by low numbers of cladocera remains, corresponding to total species densities varying between 0.5 and 516 ind cm–3 (fig. 3). the no n c om me rci al us e o nly 136 j. korponai et al. fig. 3. cladocera statigraphy for marótzugi-holt-tisza oxbow. bosm_long: bosmina longirostris, alon_nana: alonella nana, grap_test: graptoleberis testudinaria, acr_harp: acroperus harpae, alon_exig: alonella exigua, alon_cost: alona costata, sida_crys: sida crystallina, alon_affi: alona affinis, pleu_laev: pleuroxus laevis, alon_quad: alona quadrangularis, pleur_trig: pleuroxus trigonellus, eur_lame: eurycercus lamellatus, disp_rost: disparalona rostrata, alon_gutt: alona guttata, leyd_leyd: leydigia leydigi, chyd_spha: chydorus sphaericus, alon_rect: alona rectangula, alon_rust: alona rustica, ilio_sp: iliocryptus sp., kurz_lati: kurzia lattissima, mono_disp: monospilus dispar, alon_exci: alonella excisa, pleu_unci: pleuroxus uncinatus, lept_sp: leptodora kindti, pleu_adun: pleuroxus aduncus, chyd_pige: chydorus piger, simo_sp: simocephalus sp., pleu_trun: pleuroxus truncatus, leyd_acan: leydigia acanthocercoides, oxyu_tenn: oxyurella tenuicaudis, ceri_ephip: ceriodaphnia ephippia, camp_rect: camptocercus rectirostris, daph_sp.: daphnia sp., no n c om me rci al us e o nly reconstruction of flood events in an oxbow lake 137 number of cladocera species (26) in this zone was slightly lower with respect to zone 1. the most common species was again b. longispina, which accounted for an average of 80% of total cladoceran abundances. the following most abundant five species, i.e., c. sphaericus, a. rectangula, leydigia leydigi (schoedler, 1863), alona quadrangularis (o. f. müller, 1785) and a. harpae, accounted together only for 1% of cladocera, while other cladocerans species occurred with very low densities and their contribution remained below 1% (fig. 2). cl-3 zone (112-72 cm) cladoceran species richness was similar to the previous zone, as 28 species were identified from the remains. b. longirostris accounted for the highest portion (51%) of cladoceran abundance, while a. rectangula, c. sphaericus, and alonella nana (baird, 1850) were found with a slightly higher than 5% (fig. 3). cl-4 zone (72-0 cm) in the top zone, 28 cladocera species were identified, among which b. longirostris again showed an overwhelming dominance exceeding 90% of the total cladocera abundance. only c. sphaericus reached a 3% of the total cladocera abundance, while all the other species did no exceed 1% (fig. 3). simper analysis revealed that b. longirostris, c. sphaericus, a. recangula, a. harpae, l. leydigi, a. quadrangularis, and a. nana were the most important species in determining the dissimilarity of cladoceran assemblages during lotic and lentic stages in the oxbow lakes. the species number was somewhat higher in lentic layers than in lotic ones (31 and 28 respectively). remains of iliocryptus sp. (sars, 1862) and monospilus dispar (sars, 1862) were only found in lotic layers, while remains of alona rustica (scott, 1895), alonella excisa (fischer, 1854), chydorus piger (lilljeborg, 1853), kurzia latissima (kurz, 1875) and simopcephalus sp. (schoedler, 1858) occurred only in lentic ones, the others were common to both. lentic layers also contained more remains, and average cumulative density of lentic cladocerans was twice so high that of lotic ones (23,986 ind cm–3 and 11,685 ind cm–3, respectively). statistical analysis of water regime states linear discriminant analysis (lda) of subfossil cladoceran assemblages confirmed the differences between lotic and lentic water regime stages. group centroids of the lda scores of the different water regimes were well significantly separated (wilks’s λ=0.7866, df=1, p<0.05, fig. 4), and lda scores were significantly correlated with the proportion fine sand (r=-0.3695, t=-6.0434, df=231, p<0.001). lda scores were positive during lentic stages, while negative scores reflected lotic conditions. smoothed lda discriminant scores exhibited strong negative correlation with fine sand fraction (r=-0.4658, t=-8.0006, df=231, p<0.001). loi content correlated negatively with the fine sand fraction (r=-0.3288, t=-5.2918, df=231, p<0.001), and positively with lda scores (r=0.3169, t=5.0795, df=231, p<0.001). furthermore, spdu exhibited a weak negative correlation to the fine sand fraction (r=-0.1622, t=-2.4996, df=231, p<0.013), as well as a positive relation with discriminant scores (r=0.3168, t=5.0768, df=231, p<0.001). lotic sediment layers showed significantly lower loi and spdu values (loi: k-w test, χ2=9.9031, df=1, p<0.01; spdu: k-w test, χ2=6.7706, df=1, p<0.01). in order to test for significant differences in cladocera community composition between lotic and lentic stages, we redefined lotic and lentic regimes by lda scores. layers with negative lda scores were considered as lotic, while layers with positive lda scores were put into lentic groups. both loi and spdu contents were significantly higher in lda based lentic sediment layers (loi: k-w test, χ2=29.165, df=1, p<0.001); spdu: k-w test, χ2=42.352, df=1, p<0.001). discussion our sediment core was 120 cm longer than the core analyzed by braun et al. (2000, 2009), and it covered a time span of 150 years, i.e., from oxbow establishment to the present. the estimated average sedimentation rate (3.53 cm y–1) agrees with the values reported by braun et al. (2009). the sediment record represents two main developmental stages of marózugi-holt-tisza. from the oxbow establishment in 1860 to 1950 the water regime of the lake was determined by height and duration of floods (fig. 2). sedimentation rates were quite high, with a mean value of 3.6 cm y–1. the construction of a large dam for electricity generation and irrigation supply for the eastern part of hungary was started in 1950 and concluded in 1959 at kisköre (fig. 1). the dam has increased minimum water levels in the riverbed by 3 m and its effect can be recognized upstream for 623 river km (lászlóffy, 1982). this enhanced the groundwater inflow to the oxbow, which in turn stabilized its water regime and established stable lentic conditions (babka et al., 2011). increasing organic materials and pigment contents were measured in the sediments of marózugi-holt-tisza after 1950 (fig. 2). since meroand euplanktonic algae exhibit higher production in lentic or slow running waters (istvánovics and honti, 2011), and high organic material and pigment contents are found in sediments of highly productive lakes (korponai et al., 2010a, 2011), the increased loi and spdu content has been interpreted as a marker for to the development of lentic conditions in the oxbow since 1950. size distribution of sediment sand no n c om me rci al us e o nly 138 j. korponai et al. fractions was used for tracking flood events. the bulk of the suspended solids (ss) arrive and settle in the floodplain at the beginning of a flood (csépes et al., 2002; schweitzer et al., 2002), and loading depends on the intensity of flush. in fact, coarser fractions settle and accumulate in the floodplain regions closer to the riverbed, while finer particles settle in oxbows as ss (kiss and fejes, 2000). therefore, high proportions of medium and fine sand fractions in sediment of marotzugi-holt-tisza may correspond to major flood events. using the proportions of sand fractions of sediment layers to define lotic (during floods) and lentic stages (in periods between floods) of the oxbow studied, we found significantly higher loi and spdu contents in lentic stages in the sediment profile (fig. 2). reconstructed cladocera communities of marotzugiholt-tisza oxbow consisted of species which are very common in hungary, and can be collected from almost all types of lakes and ponds. the observed shifts in cladocera communities reflected environmental changes in the oxbow. absence of cladocera remains in the deepest core layers likely indicate that this sediment section represents the original riverbed (fig. 3). zsuga (1981) found no benthic cladocerans in the upper portion of river tisza, but she found some species with very low densities in the lower river stretch. moreover, investigations of the stream complex of the lower river tisza revealed structured zooplankton communities in slow-flowing streams and in oxbows (pujin et al., 1986; pujin and ratajac, 1988; ratajac, 1989, 1992). regime shifts from lotic to lentic conditions of large rivers in hungary (river danube and river tisza) could be associated to increasing zooplankton diversity and density (vadadi-fülöp et al., 2008, 2009; zsuga, 1998). as very few cladoceran remains were found in cl-1 zone, we interpreted this section as deposited during a lotic, flood-determined stage (fig. 3). since 1950, the oxbow has become more lentic, as confirmed by the pattern in the smoothed curve of lda scores, which serves as a proxy for water regime shift. in fact, lda scores remain negative during lotic stages, while become positive during lentic stages of the marótzugi-holt-tisza (fig. 2). oxbow lakes and shallow reservoirs typically show dense macrophytes beds with high cladoceran abundances (gulyás and forró, 1992; zsuga et al., 2004, korponai et fig. 4. distribution of discriminant (lda) scores in lentic and lotic status of sediment layers. no n c om me rci al us e o nly reconstruction of flood events in an oxbow lake 139 al., 2010a, 2010b, 2010c). remains of a. harpae, e. lamellatus, disparalona rostrata (koch, 1841), pleuroxus trigonellus (o. f. müller, 1785), and sida crystallina (o. f. müller, 1776) indicate presence of macrophytes in the oxbow (gulyás and forró, 1992; korhola and rautio, 2001; korponai et al., 2010a). large floods occurred mainly in spring and autumn, while lower water levels in summer promoted the development of the macrophyte belt in marótzugi-holt-tisza. as a consequence, remains of phytophyllous cladocerans were found also in the lotic layers of the core (figs. 2 and 3). cl-1 zone covers the timespan 1876-1950, that can be described as the period of regulation works at river tisza, when subsequent large floods inundated its floodplains covered them with thick sediment layers. during floods thick suspended solids layers buried all habitats that prevented recovery of cladoceran community from egg banks (vaničková et al., 2011), but resting eggs could been washed in with floods from other habitats allowing development of new populations after floods (havel et al., 2000). the low number of remains in this core section therefore can be explained by the slow recovery of the cladoceran community from sediment egg banks. in cl-2 and cl-3 zones (between 1950-1970) the co-occurrence of a high proportion of bosmina longirostris and of phytophyllous cladoceran species indicates the development of a macrophyte bed with large open water patches due to a lentic water regime (galbarczyk-gąsiorowska et al., 2009). in cl-4 zone, the absolute dominance of b. longirostris since 1970 is related to strong human impacts. in fact, small bodied cladocerans as b. longirostris typically dominate in eutrophic waters with high fish abundance (jeppesen et al., 2011; goslar et al.,1999; galbarczyk-gąsiorowska et al., 2009). the high proportion of b. longirostris in marótzugi-holt-tisza can be related to a fish effect, since this oxbow has been utilized as a fish pond since 1961. the majority of species composing the cladoceran assemblages found in the sediment studied were common throughout the whole core. we did not find remarkable differences in species compositions of lotic and lentic communities, but densities were higher during lentic conditions. three-fold more remains were found in lentic layers than in lotic ones. remains of mud dwellers, i.e., iliocryptus sp. and monospilus dispar, were found exclusively in lotic sediment layers. these species prefer sandy or muddy surfaces, which establish after floods (sebestyén, 1965, 1970, frey, 1986, 1988, kattel et al., 2007). species which were found in lentic layers of marótzugi-holt-tisza (a. rustica, a. excisa, c. piger, k. latissima and simocephalus sp.) were recorded from shallow lakes and ponds in hungary in macrophytes belts (korponai et al., 2010a). galbarczyk-gąsiorowska et al. (2009) found that remains of a. excisa, and k. latissima were associated with macrophyte occurrence in a lakebog transition zone in stare biele mire. although a large number of common species were found both in lotic and lentic layers, lda analysis confirmed separation of the two stages (fig. 4). for testing how cladocerans can indicate lotic/lentic states, we redefined lotic and lentic attributes of sediment layers by lda scores. the two lda-based regime were clearly different from each other, thus outlining that lotic and lentic water regimes can be distinguished based on their cladoceran communities . in order to understand how much ldabased lotic-lentic categories correspond to proportion of fine sand, we compared how many layers changed their attributes. we found that 71% of sediment layers kept their original status, 14% changed their own character from lotic to lentic, while 15% of layers turned from lentic to lotic. this means that if we determine lotic or lentic conditions based only on subfossil cladocera communities, our determination would be correct in 71% of cases. conclusions marótzugi-holt-tisza has gradually evolved from a lotic to a lentic (pond) system during the last ~150 years, and all the stages of this evolution were recorded in sediment studied. marótzugi-holt-tisza was lotic until 1950 but has now become a lentic system. this transition was determined by the frequent floods of river tisza and it was reflected by changes in the zooplankton community. lotic systems showed lower zooplankton densities, but were typically characterized by mud dweller species, as monospilus dispar and iliocryptus sp. bosmina longirostris can be regarded as indicator of human impact, since the studied oxbow has been intensively utilized as a fishpond for anglers. this work showed that flood events can drive changes in cladoceran communities, and that subfossil cladocera remains can be successfully used to track shifts from lotic to lentic water regimes in oxbow lakes. acknowledgments this study was financially supported by the hungarian national science foundation, otka-t 049098 and the hungarian national research and development program balöko 3b022/04, támop 4.2.2-08/1-2008-0020, támop 4.2.1/b-09/1/konv-2010-0006. references babka b, futó i, szabó s, 2011. clustering oxbow lakes in the upper-tisza region on the basis of stable isotope measurements. j. hydrol. 410:105-113. no n c om me rci al us e o nly 140 j. korponai et al. botár i, károlyi zs, 1971. [the regulation of the tisza. part i. (1846-1879)].[article in hungarian]. vizdok, budapest. braun m, papp i, korponai j, lukács v, gyulai i, forró l, hubay k, szalóki i, 2010. [traces of water level changes in the sediment of marótzug oxbow lake].[article in hungarian with english abstract]. hidrol. közl. 90:20-22. braun m, tóth a, alapi k, dévai g, lakatos g, posta j, szalóki i, 2000. environmental history of oxbow ponds: a sediment geochemical study of marót-zugi-holt-tisza. tiscia 5:133-138. csépes e, bancsi i, végvári p, aranyné rózsavári a, 2003. [investigation of alluvion in the middle section of river tisza (from kiköre to szolnok)].[article in hungarian]. proc. 21st ann. meet. hungarian hydrological society 2-3:1-10. frey dg, 1950. the taxonomic and phylogenetic significance of the head pores of the chydoridae (cladocera). int. rev. ges. hydrobiol. 44:27-50. frey dg, 1962. cladocera from the eemian interglacial of denmark. j. paleontol. 36:1133-1154. frey dg, 1986. cladocera analysis, p. 692-667. in: b.e. berglund (ed.), handbook of palaeooecology and palaeohydrology. j. wiley & sons, chichester. frey dg, 1988. littoral and offshore communities of diatoms, cladocerans and dipterous larvae, and their interpretation in paleolimnology. j. paleolimnol. 1:179-191. frey dg, 1991. first subfossil records of daphnia headshields and shells (anomopoda, daphniidae) about 10 000 years old from northernmost greenland, plus alona guttata (chydoridae). j. paleolimnol. 6:193-197. galbarczyk-gąsiorowska l, gąsiorowski m, szeroczyńska k, 2009. reconstruction of human influence during the last two centuries on two small oxbow lakes near warsaw (poland). hydrobiologia 631:173-183. goslar t, ralska-jasiewiczowa m, van geel łącka b, szeroczyńska k, chróst l, walnus a, 1999. anthropogenic changes in the sediment composition of lake gościąż (central poland), during the last 330 yrs. j. paleolimnol. 22:171-185. goulden ce, frey dg, 1963. the occurrence and significance of lateral head pores in the genus bosmina (cladocera). int. rev. ges. hydrobiol. 48:513-522. grimm ec, 1987. coniss: a fortran 77 program for stratigraphically constrained cluster analysis by the method of incremental sum of squares. comput. geosci. 13:13-35. gulyás p, forró l, 1999. [a guide for the identification of cladocera occurring in hungary].[book in hungarian]. kgi, budapest. gulyás p, forró l, 1992. composition and abundance of microcrustacean fauna in the upper reservoir (hídvégi-tó) of the kis-balaton. miscnea zool. hung. 7:39-51. havel je, eisenbacher em, black aa, 2000. diversity of crustacean zooplankton in riparian wetlands: colonization and egg banks. aquat ecol. 34:63-76. heiri o, lotter af, lemcke g, 2001. loss on ignition as a method for estimating organic and carbonate content in sediments: reproducibility and comparability of results. j. paleolimnol. 25:101-110. istvánovics v, honti m, 2011. phytoplankton growth in three rivers: the role of meroplankton and the benthic retention hypothesis. limnol. oceanogr. 56:1439-1452. jeppesen e, nõges p, davidson ta, haberman j, nõges t, blank k, lauridsen tl, søndergaard m, sayer c, laugaste r, johansson ls, bjerring r, amsinck sl, 2011. zooplankton as indicators in lakes: a scientific-based plea for including zooplankton in the ecological quality assessment of lakes according to the european water framework directive (wfd). hydrobiologia 676:279-297. juggins s, 2009. rioja: analysis of quaternary science data, r package ver. 0.5-6. available from: http://cran.r-project.org/ package=rioja. kattel gr, battarbee rw, mackay a, birks hjb, 2007. are cladoceran fossils in lake sediment samples a biased reflection of the communities from which they are derived? j.paleolimnol. 38:157-181. kiss t, fejes a, 2000. flood caused sedimentation on the foreshore of the river tisza. acta geograph. 37:51-55. knowlton me, jones jr, 1997. trophic structure of misouri river floodplain lakes in relation to basin type and connectivity. wetlands 17:468-475. korhola a, rautio m, 2001. 2. cladocera and other branchiopod crustaceans, p. 1-38. in: j.p. smol, h.j.b. birks and w. m. last (eds.), tracking environmental change using lake sediments, vol. 4. zoological indicators. kluwer academic publishers, dordrecht. korponai j, braun m, buczkó k, gyulai i, forró l, nédli j, papp i, 2010a. transition from shallow lake to a wetland: a multi-proxy case study in zalavári pond, lake balaton, hungary. hydrobiologia 641:225-244. korponai j, braun m, gyulai i, forró l, nédli j, papp i, 2010b. [are the cladocera remains in sediment suitable for reconstruction of diversity of cladoceran community of lakes?]. [article in hungarian with english abstract]. hidrol. közl. 90:66-68. korponai j, braun m, gyulai i, forró l, nédli j, papp i, 2010c. [paleolimnological reconstruction of naturally cut-off oxbow by geochemical and cladocera remains proxies].[article in hungarian with english abstract]. hidrol. közl. 90:68-70. korponai j, k. varga ka, lengré t, papp i, tóth a, braun m, 2011. paleolimnological reconstruction of the trophic state in lake balaton (hungary) using cladocera remains. hydrobiologia 676:237-248. lászlóffy w, 1982. [the river tisza].[book in hungarian]. akadémiai kiadó, budapest: 610 pp. luoto tp, nevalainen l, kultti s, sarmaja-korjonen k, 2011: an evaluation of the influence of water depth and river inflow on quantitative cladocera-based temperature and lake level inferences in a shallow boreal lake. hydrobiologia 676:143-154. müller z, dévai gy, miskolczi m, kiss b, tóth a, nagy s, grigorszky i, jakab t, 2000. [study on dragonflies as indicators of biotope heterogeneity in the active floodplain of river tisza between tiszabercel and gávavencsellő].[article in hungarian with english abstract]. hidrol. közl. 56:373-376. nyárádi l, 1900. [water regulation and floodprotection, p. 278294]. in: s. borovszky and j. sziklay (eds.), [counties and cities of hungarian kingdom].[book in hungarian]. budapest. oksanen j, blanchet gj, friendly m, kindt r, legendre p, mcglinn d, minchin pr, o’hara pb, simpson gl, solymos no n c om me rci al us e o nly reconstruction of flood events in an oxbow lake 141 p, stevens mhh, szoecs e, wagner h, 2016. vegan: community ecology package. r package version 2.4-1. available from: https://cran.r-project.org/package=vegan. pálfai i, 2001. [oxbow-lakes in hungary].[in hungarian with english abstract]. ministry for transport and water management: budapest, budapest. pujin v, ratajac r, 1988. structure and dynamics of zooplanktonin the dead theiss. tiscia 23:51-59. pujin v, ratajac r, djukić n, 1986. [ein beitrag zur limnologischen untersuchungen der carska bara].[article in german]. tiscia 21:69-80. r development core team, 2010. r: a language and environment for statistical computing. r foundation for statistical computing, vienna, austria. available from: http://www.rproject.org ratajac r, 1989. the composition and the dynamics in population of the dominant crustacea species in mrtva tisza. tiscia 24:49-57. ratajac r, 1992. the structure and dynamics of cladocera in the yugoslavian section of the river tisza. tiscia 26:59-61. schweitzer f, nagy i, alföldi l, 2002. [relationship between the formation of point bars and natural levees and flood bed sedimentation along the middle stretches of tisza river].[article in hungarian]. föld. ért. 51:257-278. sebestyén o, 1965. [cladocera studies in lake balaton iii. preliminary studies for lake history investigations].[article in hungarian with english abstract]. annal. biol. tihany 36:229-256. sebestyén o, 1969.[cladocera studies in lake balaton iv. quaternary remains in the sediment of lake balaton i].[article in hungarian with english abstract]. annal. biol. tihany. 36:229-256. sebestyén o, 1970. [cladocera studies in lake balaton iv. quaternary remains in the sediment of lake balaton ii].[article in hungarian with english abstract]. annal. biol. tihany. 37:247-279.). sebestyén o, 1971. [cladocera studies in lake balaton iv. quaternary remains in the sediment of lake balaton iii].[article in hungarian with english abstract]. annal. biol. tihany. 38:227-268. szeroczyńska k, sarmaja-korjonen k, 2007. atlas of subfossil cladocera from central and northern europe. friends of the lower vistula society, świecie, poland. vadadi-fülöp cs, 2009. zooplankton (cladocera, copepoda) dynamics in the river danube upstream and downstream of budapest, hungary. opusc. zool. budapest 40:87-98. vadadi-fülöp cs, mészáros g, jablonszky gy, hufnagel l, 2008. the zooplankton of the ráckeve-soroksár danube: spatio-temporal changes and similarity patterns. appl. ecol env. res. 6:121-148. vallentyne jr, 1955. sedimentary chlorophyll determination as a palaeobotanical method. can. j. bot. 33:304-313. vaníčková i, seda j, petrusek a, 2010. the stabilizing effect of resting egg banks of the daphnia longispina species complex for longitudinal taxon heterogeneity in long and narrow reservoirs. hydrobiologia 643:85-95. whiteside mc, williams jb, white cp, 1978. seasonal abundance and pattern of chydorid, cladocera in mud and vegetative habitats. ecology 59:1177-1188. wolfe bb, hall ri, last wm, edwards twd, english mc, karst-riddoch tl, paterson a, palmini r, 2006. reconstruction of multi-century flood histories from oxbow lake sediments, peace-athabasca delta, canada. hydrol. process. 20:4131-4153. zsuga k, 1981. benthic entomostraca fauna of the tisza and its tributaries. tiscia 16:183-190. zsuga k, 1998. spatial heterogeneity and mosaic-like structure of zooplankton in kisköre reservoir. int. rev. hydrobiol. 83:199-202. zsuga k, 1999. zooplankton investigations in the upper tisa region. tiscia monogr. ser. 4:393-399. zsuga k, tóth a, pekli j, udvari zs, 2004. [the changes in the zoopklankton of the watershed of the river tisza from 1950s to the present].[article in hungarian with english abstract]. hidrol. közl. 84:175-178. no n c om me rci al us e o nly layout 1 introduction plankton and benthos are traditionally considered as distinct communities: the first living suspended in the water column, the second in strict association with the sea-bottom. however, despite the formal distinction, plankton and benthos are strongly interconnected in coastal marine ecosystems (boero et al., 1996; griffiths et al., 2017), in virtue of: i) continuous downward fluxes of organic detritus produced by plankton and consumed at the sea bottom; ii) intermittent upward fluxes of inorganic nutrients released by benthic bacteria; and iii) periodic formation of benthic resting stages in planktonic protists and metazoans and release of planktonic larval stages by benthic animals. a further, but less investigated route for plankton-benthos coupling is represented by trophic interactions. these are based on fluxes of living matter among organisms set at different trophic levels and can be roughly categorized as either nonselective or selective feeding. nonselective feeding, involving plankton unicellular producers as food source and benthic organisms as consumers, is the most renowned of such interrelationships – e.g., the remarkably strong suspension-feeding carried out by benthic organisms in shallow coastal regions (gili and coma, 1998; lucas et al., 2016). yet, selective feeding (i.e., the active catching of living preys) is seldom reported among trophic interactions involving plankton and benthos (hoeksema and waheed, 2012). in this paper, we report direct, de visu evidence that the benthic mediterranean rainbow wrasse coris julis (linnaeus, 1758, labridae) selectively feeds on the colonial salp pegea confoederata (forskål, 1775, thaliacea). this trophic relationship was documented in the gulf of naples (gon, tyrrhenian sea, mediterranean sea, italy) in the course of a citizen science investigation employing recreational scuba-diving and carried out at the top side of banco di santa croce, an underwater rocky outcrop whose higher pinnacles set at the boundary between the benthic and pelagic realms. we present and describe photographic frames documenting the above-mentioned trophic relation, we discuss plankton-benthos coupling in light of the existence of trophic routes connecting pelagic tunicates and benthic fish and we eventually analyse conceptually some possible perturbations to this route induced by global change. methods the banco di santa croce (bsc) is a submerged rocky outcrop located 700 m off the coast of vico advances in oceanography and limnology, 2017; 8(2): 235-241 short note doi: 10.4081/aiol.2017.6973 this work is licensed under a creative commons attribution-noncommercial 4.0 international license (cc by-nc 4.0). plankton food for benthic fish: de visu evidence of trophic interaction between rainbow wrasse (coris julis) and pelagic tunicates (pegea confoederata) domenico d’alelio,1* gabriella luongo,2 iole di capua1 1stazione zoologica anton dohrn, villa comunale, 80121 naples; 2amici degli abyssi diving, via giuseppe bonito 198, 80053 castellammare di stabia, italy *corresponding author: dom.dalelio@gmail.com abstract salps (pelagic tunicates) are rarely observed in nature and trophic interactions involving them mainly rely on the inspection of stomach contents of their potential predators. moreover, salps have soft bodies that are hardly identified in potential consumers. we involved recreational scuba-divers and photographers in collecting de visu evidence of i) massive occurrence of salps and ii) trophic interactions involving salps as preys and benthic animals as consumers. direct evidence of trophic interactions between salps and benthic fish was documented by photographic frames. we detected a long colony of the salp pegea confaederata being transported by currents close to the substrate on top of banco di santa croce, an underwater rocky outcrop in the gulf of naples (italy). an individual of the rainbow wrasse coris julis attacked the above-mentioned salp colony by selectively detaching individuals and biting their stomach. our report of a trophic interaction between labrids and salps is the second in fifty years and the previous one was only indirect. in this study, citizen science allowed detecting both neglected marine animals like salps and trophic interactions involving them. visual, direct evidence of predation on salps by benthic fish adds further knowledge about patterns of living-matter fluxes between plankton and benthos, opening new questions on the potential of global change in modifying the efficient circulation of organic matter in marine systems. key words: plankton; benthos; labrids; salps; marine food-webs; coastal systems; citizen science. received: august 2017. accepted: october 2017. no nco mm er cia l u se on ly d. d’alelio et al.236 equense, on the eastern side of the gon (40° 40.68’n, 14° 26,00’e, fig. 1). bsc includes a number of pinnacles elevating towards the sea surface while its base sets at 65 m below the sea level. the observations herein presented have been gathered during a recreational scuba-diving session carried out on top of the pinnacle named secca principale (sp), at -12 m. during february 2017, ten of such diving sessions were carried out. this period of observation was chosen based on the fact that the yearly maximum abundance of salps in the gon was recorded in winter and summer (mazzocchi et al., 2011). pictures presented herein have been taken underwater using a nikon d7100 camera hold into a nauticam housing with two sea & sea ys-110 α strobes. in addition to visual census, reference salp abundances at basin level were gathered at the long term ecological research station marechiara (lter-mc, 40°48.5’n, 14°15’e, fig. 1) (ribera d’alcalà et al., 2004; mazzocchi et al., 2011). meso-zooplakton (0.220 mm) were collected with a wp2 net (mouth area = 0.25 m2, mesh = 200 μm) towed vertically from -50 m to sea surface. samples were fixed immediately after collection and preserved in a 4% buffered formaldehydeseawater solution. once in the laboratory, samples were concentrated via filtering through a smaller mesh (<200 μm) and concentrated zooplankton samples were diluted with filtered seawater in a graduated cup up to a volume of 100-200 ml depending on sample richness. aliquots of 5 ml were then sub-sampled by stempel pipettes and analysed in a mini-bogorov chamber under the stereomicroscope at different magnifications. results during february 2017, the sp section of bsc appeared to be colonized by different species of algae, anthozoans, crustaceans and benthic fish, among which the two labrids species corius julis and thalassoma pavo (linnaeus, 1758). colonies of salps were also regularly detected and pegea confoederata (fig. 2) was one of the most recorded species. all along the same period, the average salp abundance at lter-mc, which is almost 11 fig. 1. the gulf of naples (tyrrhenian sea, mediterranean sea). geographic localization of banco di santa croce (bsc) and long term ecological research station marechiara (lter-mc). no nco mm er cia l u se on ly salp-labrid trophic interaction 237 nautical miles far from bsc, accounted for 5.6±1.4 individuals per m–3. nonetheless, p. confaederata was not detected in plankton samples collected at lter-mc. a colony of p. confaederata was floating in the open water in the immediacy of the bsc on 18 february 2017. this colony was transported by currents nearby the sp substrate and the following trophic interaction quickly established (fig. 3 a-h): i) a solitary rainbow wrasse moved towards the salp colony and attacked the exterior blastozoids; ii) the fish initially bit salp’s stomach, while detaching its prey from the colony; iii) in few minutes, the fish ate the interior part of the blastozoid and left the gelatinous body of the salp torn and empty; iv) the fish attacked the colony again, biting the stomach of a second blastozoid and making the latter detaching from the colony; v) the fish ate specifically the internal part of the salp at the second attack too. as soon as water-current slowly brought p. confoederata colony away from the proximity of the substrate and towards the open waters, the wrasse stopped eating and did not follow the chain of blastozoides. remarkably, the wrasse did not bite pelagic crustaceans setting in the interior side of the salp body, i.e., the ovigerous females of copepod sapphirina spp. (thompson, 1830) (fig. 2 b-c) and some specimens of amphipod hyperiidea (milne-edwards, 1830) (fig. 2 a-b). discussion coris julis is an omnivore fish with a preference for animal material (karachle and stergiou, 2017), it predates mainly benthic gastropods and crustaceans but even ectoparasites borne by fishes of same size (e.g., boobs boops) or even larger (e.g., mola mola) than labrids (bertoncini et al., 2009; moosleitner, 1980; narvaez et al., 2015; vasco-rodrigues and cabrera, 2015). when fig. 2. the salp pegea confaederata at banco di santa croce. a) closer view of a colony of blastozoids, the sexual life-cycle stages. b) closer view of salp’s stomach (i.e., the orange circular organ at the right side of the picture) and of female copepods of genus sapphirina (whitish animals) and amphipods (brownish animals) adhering to salp’s body. c) closer view of salp’s ‘tunic’ (i.e., the gelatinous and cellulose-based body envelop) and of parasitic sapphirina spp. no nco mm er cia l u se on ly d. d’alelio et al.238 fig. 3. labrid predation on salps at banco di santa croce. a) long colony of pegea confaederata blastozooids. b-h) different phases of the predatory attack made by an individual of coris julis on a colony of the salp; b) the fish removes a single blastozoid from the salp colony; c-d) the fish eats selectively the stomach of the salp individual; e-f) the fish attacks the prey following the same modality as above; h) the fish does not eat the gelatinous envelop of the salp. no nco mm er cia l u se on ly salp-labrid trophic interaction 239 behaving as a cleaner, c. julis is an opportunist feeder searching for preys attached to swimming hosts, which are inspected by means of high frequency bites (narvaez et al., 2015). this consideration let us hypothesize that the trophic interaction reported herein arose from the visual attraction exerted by the slow moving and macroscopic salp-colony on the wrasse. this latter was plausibly attracted by the coloured (and nutritive) stomach placed at the lowest portion of the salp body. we cannot exclude that c. julis confused the salp stomach with an ectoparasite. salps are important components of the diet of 149 species of fishes (a review is presented in henschke et al., 2016). nonetheless, only one labrid species was reported to predate salps to date, i.e., the tropical wrasse clepticus parrae, whose stomach included mainly plankton, and salps accounted for 4.7% of total ingested food (randall, 1967). there is no report of predation on salps by mediterranean-wrasses to date, but salps are soft-bodied animals, they are digested rapidly and their identification by means of interior inspection of potential predators is weakly effective. salps represent good targets to intermediate predators, including small fish, for a number of reasons: i) salps are macroscopic and easily detectable; ii) while feeding, they slowly move thanks to pulse movements of their gelatinous body and this may favour encounter with predators (bone, 1998); iii) salps’ energetic value is relatively high (henschke et al., 2013), since they concentrate a big amount of organic material via filter feeding, with preference on smaller and more nutritive plankton particles (sutherland et al., 2010). salps are an apparently infrequent food for fish, probably because they have patchy geographic distributions and aperiodic population growths. the major demographic increases of these gelatinous organisms usually coincide with sudden and short-lived swarms of long colonies (up to 20 m in length) formed by specific life-cycle stages called blastozoides (boero et al., 2013). the latter reproduce sexually and alternate with the solitary form called oozoid, which reproduces asexually via the release of colonial aggregates of blastozoids. classical mesozooplankton sampling activity (e.g., wp2 net with vertical haul operated from research vessels) are not suitable for detecting salp outbreaks since in many cases samplers destroy colonies. lack of systematic observations of salp populations is probably at the base of the dispute about the actual increase of these organisms in the present warming oceans (see condon et al., 2012). nonetheless, salps may be favoured in the future oceans by: i) the positive effect of temperature on their feeding and reproductive performances (bone, 1998); ii) their potentially high physiological adaptation to lower phytoplankton concentrations (deibel and paffenhofer, 2009), and the latter are significantly regressing in the global ocean since the beginning of last century (boyce et al., 2010); iii) their higher affinity to picoand femtoplankton particles (sutherland et al., 2010), whose relative dominance is thought to be expanding in warmer oceans (peter and sommer, 2013), in consequence of reduced nutrient availability in the photic zone (chust et al., 2014). in consequence of global warming, mediterranean wrasses are experimenting changes in latitudinal and bathymetric distributions of their populations (milazzo et al. 2011, 2016), and such changes can synergistically interact with salps’ increase and affect marine food-web structure and dynamics. coris julis is a cold-adapted species and water-temperature rising will induce in this fish both weaker physiological performances and shifting of feeding activity towards epibenthic organisms (milazzo et al., 2013); indeed, these latter lay on a still substrate (contrarily to ectoparasites, most of them laying on moving animals) and can be more easily caught. the possible demographic regression of c. julis populations may be flanked by increasing relative dominance of the sympatric t. pavo (milazzo et al., 2013), which is reported to feed on a less diverse array of preys (narvaez et al., 2015) and it is apparently less attracted by salps (according with the present study). in a context of water-temperature rising, the present winner (the warm-adapted t. pavo) is not necessarily a ‘fitter’ winner, because it is apparently less prone to adapt its trophic behaviour to relatively more available resources, such as salps. benthic ecosystems sustain coastal fisheries (kritzer et al., 2016) and are also severely impacted by global change (harley et al., 2006), with temperature rising inducing major ecological detriments (e.g., rivetti et al., 2014). even if plankton differ in many aspects from benthos, the interaction between these dominions can impact coastal food webs (boero et al., 1996; griffiths et al., 2017); to this respect, interaction with benthos should be integrated within conceptual and computational models of plankton food-webs, which are still at their infancy (e.g., d’alelio et al., 2016a, 2016b). conclusions in conclusion, underwater rocky crops are promising study-sites for plankton-benthos coupling: they are highly productive sites (pitcher et al., 2008), represent ideal points of junction between planktonic and benthic communities, provide suitable conditions for the establishment of trophic interactions among organisms that are mainly associated to one of these sub-systems and, ultimately, they are relatively easily accessible to scuba-divers. to the latter respect, observational studies supported by citizen scientists are particularly promising. no nco mm er cia l u se on ly d. d’alelio et al.240 acknowledgements dda and idc thank the flagship project ritmare the italian research for the sea coordinated by the italian national research council and funded by the italian ministry of education, university and research within the national research program 2011-2013 for providing funds. this work is dedicated to the memory of ugo di capua, a pioneering scuba diver at banco di santa croce. conflict of interest the authors declare no competing interests. ethical statement this article does not contain any studies with human participants or animals performed by any of the authors. data included herein have been collected in nature as photographic material of living animals in their specific environment. references bertoncini áa, machado lf, barreiros jp, hostim-silva m, verani jr, 2009. cleaning activity among labridae in the azores: the rainbow wrasse coris julis and the azorean blue wrasse centrolabrus caeruleus. j. mar. biol. assoc. uk 89:859-861. boero f, belmonte g, bracale r, fraschetti s, piraino s, zampardi s, 2013. a salp bloom (tunicata, thaliacea) along the apulian coast and in the otranto channel between marchmay 2013. f1000research 2:181. boero f, belmonte g, fanelli g, piraino s, rubino f, 1996. the continuity of living matter and the discontinuities of its constituents: do plankton and benthos really exist? trends ecol. evol. 11:177-180. bone q, 1998. the biology of pelagic tunicates. oxford university press, oxford: 362 pp. boyce dg, lewis mr, worm b, 2010. global phytoplankton decline over the past century. nature 466:591-596. chust g, allen ji, bopp l, schrum c, holt j, tsiaras k, zavatarelli m, chifflet m, cannaby h, dadou i, daewel u, wakelin sl, machu e, pushpadas d, butenschon m, artioli y, petihakis g, smith c, garçon v, goubanova k, le vu b, fach ba, salihoglu b, clementi e, irigoien x, 2014. biomass changes and trophic amplification of plankton in a warmer ocean. glob. chang. biol. 20:2124-2139. condon rh, graham wm, duarte cm, pitt ka, lucas ch, haddock shd, sutherland kr, robinson kl, dawson mn, beth m, mills ce, purcell je, malej a, mianzan h, uye s, 2012. questioning the rise of gelatinous zooplankton in the world’s oceans. bioscience 62:160-169. d’alelio d, libralato s, wyatt t, ribera d’alcalà m, 2016a. ecological-network models link diversity, structure and function in the plankton food-web. sci. rep. 6:21806. d’alelio d, montresor m, mazzocchi mg, margiotta f, sarno d, ribera d’alcalà m, 2016b. plankton food-webs: to what extent can they be simplified? adv. oceanogr. limnol. 7:5646. deibel d, paffenhofer g-a, 2009. predictability of patches of neritic salps and doliolids (tunicata, thaliacea). j. plankton res. 31:1571-1579. gili jm, coma r, 1998. benthic suspension feeders: their paramount role in littoral marine food webs. trends ecol. evol. 13:316-321. griffiths jr, kadin m, nascimento fja, tamelander t, törnroos a, bonaglia s, bonsdorff e, brüchert v, gårdmark a, järnström m, kotta j, lindegren m, nordström mc, norkko a, olsson j, weigel b, žydelis r, blenckner t, niiranen s, winder m, 2017. the importance of benthicpelagic coupling for marine ecosystem functioning in a changing world. glob. chang. biol. 23:2179-2196. harley cdg, hughes ar, hultgren km, miner bg, sorte cjb, thornber cs, rodriguez lf, tomanek l, williams sl, 2006. the impacts of climate change in coastal marine systems. ecol. lett. 9:228-241. henschke n, bowden da, everett jd, holmes sp, kloser rj, lee rw, suthers im, 2013. salp-falls in the tasman sea: a major food input to deep-sea benthos. mar. ecol. prog. ser. 491:165-175. henschke n, everett jd, richardson aj, suthers im, 2016. rethinking the role of salps in the ocean. trends ecol. evol. 31:720-733. hoeksema bw, waheed z, 2012. it pays to have a big mouth: mushroom corals ingesting salps at northwest borneo. mar. biodivers. 42:297-302. karachle pk, stergiou ki, 2017. an update on the feeding habits of fish in the mediterranean sea (2002-2015). mediterr. mar. sci. 18:43-52. kritzer jp, delucia mb, greene e, shumway c, topolski mf, thomas-blate j, chiarella la, davy kb, smith k, 2016. the importance of benthic habitats for coastal fisheries. bioscience 66:274-284. lucas lv, cloern je, thompson jk, stacey mt, koseff jr, 2016. bivalve grazing can shape phytoplankton communities. front. mar. sci. 3:1-17. mazzocchi mg, licandro p, dubroca l, di capua i, saggiomo v, 2011. zooplankton associations in a mediterranean longterm time-series. j. plankton res. 33:1163-1181. milazzo m, mirto s, domenici p, gristina m, 2013. climate change exacerbates interspecific interactions in sympatric coastal fishes. j. anim. ecol. 82:468-477. milazzo m, palmeri a, falcón jm, badalamenti f, garcìacharton ja, sinopoli m, chemello r, brito a, 2011. vertical distribution of two sympatric labrid fishes in the western mediterranean and eastern atlantic rocky subtidal: local shore topography does matter. mar. ecol. 32: 521-531. milazzo m, quattrocchi f, azzurro e, palmeri a, chemello r, di franco a, guidetti p, sala e, sciandra m, badalamenti f, garcía-charton, ja, 2016. warming-related shifts in the distribution of two competing coastal wrasses. mar. environ. res. 120:55-67. moosleitner h, 1980. cleanerfish and cleanershrimps in the mediterranean. zool. anz. 205:219-240. narvaez p, furtado m, neto ai, moniz i, azevedo jmn, soares mc, 2015. temperate facultative cleaner wrasses selectively no nco mm er cia l u se on ly salp-labrid trophic interaction 241 remove ectoparasites from their client-fish in the azores. mar. ecol. prog. ser. 540:217-226. peter kh, sommer u, 2013. phytoplankton cell size reduction in response to warming mediated by nutrient limitation. plos one 8:1-6. pitcher tj, morato t, hart pjb, clark mr, haggan n, santos rs, 2008. seamounts: ecology, fisheries and conservation. wiley-blackwell, hoboken: 552 pp. randall je, 1967. food habits of reef fishes of the west indies. institute of marine sciences, university of miami coral gables: 94 pp. ribera d’alcalà m, conversano f, corato f, licandro p, mangoni o, marino d, mazzocchi mg, modigh m, montresor m, nardella m, saggiomo v, sarno d, zingone a, 2004. seasonal patterns in plankton communities in a pluriannual time series at a coastal mediterranean site (gulf of naples): an attempt to discern recurrences and trends. sci. mar. 68:65-83. rivetti i, fraschetti s, lionello p, zambianchi e, boero f, 2014. global warming and mass mortalities of benthic invertebrates in the mediterranean sea. plos one 9:1-22. sutherland kr, madin lp, stocker r, 2010. filtration of submicrometer particles by pelagic tunicates. p. natl. acad. sci. usa 107:15129-15134. vasco-rodrigues n, cabrera pm, 2015. coris julis cleaning a mola mola, a previously unreported association. cybium 39:315-316. no nco mm er cia l u se on ly layout 1 advances in oceanography and limnology, 2015; 6(1/2): 46-57 original article doi: 10.4081/aiol.2015.5439 introduction the construction of reservoirs and hydroelectric power plants (hpps) in the last century showed a huge increase, strongly influencing the hydrology of natural fluvial systems (christensen et al., 2004) and causing an extension of large flooded areas and artificial wetlands, which are excellent potential habitats for mosquitoes. while the mankind modified the natural hydrosystems, global warming favored the expansion and increasing of insect populations by changing the length of vegetation period and decreasing the severity of winter throughout europe. furthermore, climate change and the increased cargo transport facilitated the migration of arthropods, among them mosquito vectors (benedict et al., 2007; githeko et al., 2000; kovats et al., 2001). the effect of river regulation and other hydromorphological modifications such as the disturbance of wetland habitats on the temporal change of mosquito habitats has been a relatively neglected topic. cailly et al. (2011) demonstrated that the distribution of some culex and anopheles species was mainly driven by the repartition of their specific breeding habitats. the landscape determinants of mosquito larval habitats vary from species to species. for example, the primary malaria vectors in sub-saharan africa generally utilize small temporary habitats as breeding sites. different mosquito species prefer various breeding sites, e.g. aedes albopictus and aedes aegypti are typical container-breeding mosquitoes utilizing mostly artificial breeding sites created by human activities such as tires, catch basins, bins, various rubbish, and plant saucers. in contrast, the studied mosquitoes prefer or primarily use oxbow lakes and other natural shallow standing waters. it was assumed that landscape diversity highly affects the mosquito density (overgaard et al., 2003). however, since the identification of mosquito larval habitats over large geographic areas requires great efforts, several remotesensing techniques (brown et al., 2008) were developed, including drones and satellite images (mushinzimana et al., 2006). though the main phase of the river regulations occurred in europe from the 1800s to the mid-1900s, more area-based historical modeling of the effects of the river bank regulation on the potential abundance of eleven mosquito species in the river danube between hungary and slovakia attila trájer,1,2* kinga farkas-iványi,3,4 judit padisák,1,2 1university of pannonia, department of limnology, h-8200 veszprém, hungary; 2mta-pe limnoecology research group, egyetem utca 10, h-8200 veszprém, hungary; 3corvinus university of budapest, faculty of landscape architecture, villányi út 29-43, h-1118 budapest; 4danube research institute, mta centre for ecological research, karolina út 29, h-1113 budapest, hungary *corresponding author: atrajer@gmail.com abstract the construction of reservoirs and hydropower plants was accelerated in the past century due to the increasing aridity in many parts of the world. the effect of water regulations on the abundance of mosquito vectors is controversial. in this paper, the habitat preference of mosquitoes was investigated based on a 30-years long collection of the mosquito data in hungary, military maps and satellite images. three time phases of the analyzed section in the danube river were analyzed in order to characterize the impact of human influence on mosquito habitats: the semi-natural phase, the post channelization phase and the post hydropower dam state. geographical data referring to the years 1790, 1820, 1830, 1870, 1946 and 1955 were based on military maps, whereas the years 2004 and 2013 were analyzed by satellite imagery. the amoros-like eupotamon a plesiopotamon line represents an increasing gradient of habitat-suitability for mosquitoes. the habitat-preference of different mosquitoes to the amoros-classified water habitats was based on a monographic collection data. this dataset contains the collecting and trapping results from the 1960s to the early 2000s in hungary. we found that human-induced changes had prolonged impact on mosquito-suitable habitats, although the effect can be different for diverse mosquito species. the increase of the evenness of the mosquito fauna was seen since the mid-20th century, after the primary river regulation. the increasing areal extension of relatively warm and nutrient-rich water habitats had positive effects on the more rare members of the mosquito fauna, such as the potential malaria vector mosquito anopheles algeriensis according to the model results. summarizing, we found a strong, positive link between anthropogenic interventions and the mosquito diversity in water ecosystems. key words: river regulation; limnoecology; mosquito vectors. received: july 2015. accepted: november 2015. no nco mm er cia l u se on ly 47 a. trájer et al. recently serious efforts were made to constructing hpps, dams and drainage systems in the developing countries, e.g. china (dudgeon, 1995), africa and south-america (nilsson et al., 2005). there is wide consensus about the effects of river regulation and the decline of wetlands on the changing patterns of both the mosquito inhabited habitats and the occurrence of mosquito-borne diseases as it has happened in case of the upper mississippi valley (ackerknecht, 1945). however, the role of the pesticides and the insecticide treatments were not well understood. for example, malaria was a native disease in hungary until the middle of the 20th century, despite that the river minor regulations started in the 1820s, with the main part of the regulations performed between 1886 and 1896 (tőry, 1952). between 1933 and 1943, the countrywide case number of malaria was estimated in 10.000-100.000 per year (szénási et al., 2003). the application of the pesticide dichlorodiphenyltrichloroethane (ddt), coupled with serious public health interventions, finally led to the eradication of the prevalent focuses by 1956 (szénási et al., 2003; lőrincz, 1937). prior to the 1950s, in hungary anopheles atroparvus, anopheles maculipennis and anopheles messeae were the plausible potential vectors of the plasmodium parasites (szénási et al. 2003). despite the river regulation, mosquito-borne diseases other than malaria are emerging diseases in hungary. west nile fever and dirofilariasis caused by dirofilaria repens and dirofilaria immitis became endemic over the last two decades in the carpathian basin (fok, 2007, krisztalovics et al., 2008). additionally, aedes mosquitoes, the primary vectors of dengue and chikungunya fever, are widely present in the mediterranean basin (spain, france, italy, albania, greece, bosnia and herzegovina) and the southern neighboring countries of hungary, such as slovenia, croatia and serbia (klobučar et al., 2006; scholte and schaffner, 2007). climate change can facilitate the spread of vector aedes species in the carpathian basin and many other parts of europe. according to regional climate models, the climate of the carpathian basin in the 2071-2100 period is expected to become wetter and milder in winter season (bartholy et al., 2007, 2008) favoring the colonization of invasive mosquito species. mosquitoes are strongly related to the river bank systems and such a simple geographic factor as the distance from the river may determine the risk patterns of the acquiring mosquitoborne diseases, as in the case of the west nile fever in 2008 documented in hungary (trájer et al., 2014). since the adult female mosquitoes have good flying ability, the expansion of the mosquito species can be relatively rapid. mosquitoes in general require silent, relatively warm environments for their reproduction and development. these conditions are not those typically found in the eupotamon streams of the temperate areas of europe. the conditions of the parapotamon and plesiopotamon streams are more suitable for mosquitoes. it remains an open question whether the river regulation facilitated or impaired the habitat templates of the potential mosquito vectors. since the construction of water reservoirs and hydroelectric power plants is still increasing and the naturalization of the river banks has also started, it is an urgent task to study the potential effect of these changes and evaluate the possible effects on the mosquito populations like anopheles maculipennis or anopheles atroparvus. in western hungary, prior to extensive regulations, the szigetköz floodplain was a periodically restructuring alluvial anabranching channel system (guti et al., 2010). the rejuvenation of the vegetation was twice as intensive in its upper section as in its lower area (farkas-iványi and guti, 2013). the first major, but indirect human influence affected the area after 1812 (hohensinner et al., 2004, 2005, 2011). with the beginning of the regulation of the austrian part of the danube, vast sediment flux reached the alluvial plain of the szigetköz part of the danube, lifting the river bed. the second, direct anthropogenic effect was the confining of the danube into dams which reached its peak with the formation of the artificial main channel at the end of the 19th century. the third human influence was represented by the building of the gabčíkovo-nagymaros dams (19771992), which caused the diversion of the 90% of water off the old riverbed (fürst, 2006). the water replacement system was completed by 1995 however, in the szigetköz, the main channel received only the 15% of the river water supply, causing a drastic decrease in the water level. the modified hydro-geomorphological processes lead to terrestrialization, formation of abandoned channels and significant decline of ecological rejuvenation in the floodplain (farkas-iványi and guti, 2014). the main objective of this work is to assess the consequences of the river regulations and construction of water power plants on the abundance of mosquito vectors in the danube river, in hungary. the long-term changes will be evaluated by comparing the original function and changing in the szigetköz territory by using historical maps and reconstructing the change of the relative number of mosquito larvae during the last 200 years. the adopted method was based on the weighting of the past and present aquatic habitats with the relative habitat-preference value of the studied mosquitoes. methods according to the above described history, three time phases of river regulation were analyzed in order to categorize the impact of human influence on the river habitat and mosquito changes. the three phases include the seminatural phase (maps 1790, 1820, 1830, and 1870), the post channelization phase (maps 1946, 1955 and 1986) and the post hydropower dam state (2004, 2013). no nco mm er cia l u se on ly 48impact of river bank regulation on mosquito vectors the studied mosquito species different anopheles, aedes and culex mosquito species with different relative countrywide dominance/ frequency values (d) according to the dataset of tóth (2004) were considered in this work. these include: anopheles algeriensis theobald 1903 (d<0.1%), anopheles atroparvus van thiel 1927 (d=0.11%), anopheles claviger meigen 1804 (d=1.48%), anopheles hyrcanus pallas 1771 (d=0.03%), anopheles maculipennis meigen 1818 (d=2.85%), anopheles messeae falleroni 1926 (d=0.38%), aedes cinereus meigen 1818, aedes rossicus dolbeskin and gorickaja 1930 (d=0.57%), aedes vexans meigen 1830 (d=15.12%), culex modestus ficalbi 1890 (d=3.26%) and culex pipiens pipiens linnaeus 1758 (d=15.15%). each modeled mosquito species are native in the studied area. an. atroparvus, an. hyrcanus and an. algeriensis are thought to be potential vector of plasmodium species in the mediterranean (jetten and takken, 1994); an. maculipennis and an. messeae are also important malaria-vectors. culex pipiens is one of the potential vectors of dirofilaria immitis and avian malaria in hungary (zittra et al., 2015). definitions of the aquatic habitat types in the szigetköz floodplain in this study, 28 historical maps provided information on the change of the river morphology and the historical areas of the floodplain basin of szigetköz. the georeferencing of the historical maps of the area prepared in 1790, 1820, 1825, 1830, 1870, 1946, 1955, 1986, 2004 and 2013 was performed in arcgis 10.0, by selected landmarks, both historical and still existing. second polynomial (affine) transformation was used between the temporarily different raster layers (cajthaml, 2013).the typology of the habitat analyses was based on the functional classification by amoros et al. (1987). the amoros-classification takes into account the flow velocity of the branches and the intensity of lateral connectivity between the main channel side arms and backwaters. we selected this topological classification system because it provides a single morphological identification of the historical stages allowing a comparison of the recent state without requiring the knowledge of the long-term changes in water chemistry and other factors. changes in the water flow and lateral connectivity from the main channel (the eupotamon a) to the paleopotamon stage are in good correspondence with the main requirements of mosquito species. the determined branches were signed by consequent coloring. we calculated the area of the georeferred and colored branches (fig. 1). with the construction of the gabčíkovo-nagymaros waterworks, water replacement system cascade-like sections were included into the former main branches (mainly replacing the earlier eupotamon b). these modifications increased the water level, creating a system of numerous reservoirs. in this paper, we refer to these branches as cascadepotama. since these branches have a permanent, but relatively low flow, the corresponding habitats can be classified between the plesiopotamon and eupotamon b, with low water level. in the successive analysis we will handle the cascadepotamon as the transient stage between the two habitats, but it is important to note that this assumption is made only as a practical approximation. finally, the number of the transition numbers of the habitats of given potama (e.g., paleopotamon in 1825 to plesiopotamon in 1830) were counted between each studied neighboring years. the numbers of transitions were normalized according to the length of the periods between the adjacent years. we assumed that the collection record and long-term evolution of the different water types can reflect the effects of the different physical and chemical conditions, as well as the level of the predatory pressure to the larvae of mosquitoes. this was referred to as the relative value of mosquito suitability. further, the surface of a given water body was assumed to be linked to the potential number of larvae. larvae, breathing through spiracles or a siphon, come to the surface frequently and spend most of their time feeding on the surface microlayer of the stems and leaves of the submerged vegetation and floating-leaf plants. the characterization of the habitat-preference of the different mosquitoes was based on the countrywide monographic collection data of the fauna of mosquitoes in hungary (diptera: culicidae) by tóth (2004). this faunistic monograph contains the collection data of the breeding sites, the seasonality and the habitat-preference data of the native mosquitoes of hungary based on investigations carried out since the 1960s (cf. tab. 1). the monograph separately lists the collecting data of larvae, pupae, and the adult male and female mosquitoes. the habitats were associated with the number of breeding sites where the larvae of a given mosquito species were collected. the monograph contains the collection data of 198066 larvae of 48 different mosquito species. the indicated countrywide dominance/frequency values of the mosquito larvae (d) were also gained from tóth (2004). mosquito larvae were captured by net collection at 1) the waterline; 2) from the surface of submerged stems and shoots; and 3) from the sediment (tóth, 2004). the data regarding the larvae of mosquitoes were used according to the following standpoints in the characterization of the water bodies: 1) larvae are the most water-bound and flightless ontogeny forms of mosquitoes; compared to adults, they are easier to locate. 2) the collection of larvae is the less selective proceno nco mm er cia l u se on ly 49 a. trájer et al. dure; the selective attractiveness of traps or human bodies does not play any role in the case of larvae, ensuring a good represent activity in contrast to imagoes. 3) we considered the number of the breeding sites in the different aquatic habitats as an indicator of the habitat preference of the studied mosquito species. the weight (w) of the different aquatic habitat types was derived from the number of the breeding habitats in the countrywide summary of the mosquito occurrences. it can be assumed that if a mosquito species (a) was more frequently collected in a habitat type than another (b) species, then a prefers more that given habitat than b. the w is coincident with the values of tab. 2. floodplain and floodplain pool, the medium-sized river, the small lake reservoir, the types of artificial small lakes of standing waters, the other types of small artificial lakes of standing waters, the littoral shallow lakes, the litoprofundal shallow lakes, the artificial swamps and standing waters, the marsh-type natural standing waters, the natural or artificial duckweed covered puddles, swamps and small marshes can be classified with the amoros-scheme of water classification (tab. 3. tömpöly’s the special hungarian name of a habitat type which is defined as a minor permanent water body derived e.g. from the pits left behind of the constructions of dams (small, shallow water basins vegetated by aquatic submersed or floating macrophytes, which are perennial and do not dry out regularly). the correspondence between the aquatic habitats classified by using the amoros-like classification and the mosquito-collecting habitats identified by tóth (2004) (tab. 3) was somewhat problematic. while the first is a regional scale system mainly based on topographical characters, the latter is a subsidiary classification system based on the habitats colonized by the mosquito larvae. some of the aquatic habitats classified by amoros et al. (1987) do not occur in the studied area (e.g., cold springs) or are not quantifiable on an aerial basis (e.g., the technotelfig. 1. the spatial subdivision of aquatic habitats in the studied area in 1946 and 1955. light green field marks the floodplain forest; the white background shows the arable lands. dark blue, eupotamon a; blue, eupotamon b; light blue, parapotamon a; pale blue, parapotamon b; neon green, plesiopotamon; dark green, paleopotamon. no nco mm er cia l u se on ly 50impact of river bank regulation on mosquito vectors mata/phytotelmata habitats, which would be important, but the quantification has yet been impossible). although paleopotama and plesiopotama are isolated water bodies with only occasional connection with the main channel, paleopotama are shallower habitats mainly with a single littoral part, whereas plesiopotama have a profundal riverbed part. these differences were also reflected in the data collected by tóth (2004). it is important to highlight that the eupotamon-parapotamon-plesiopotamon-paleopotamon sequence coincides with the progression of the river bank succession, and with the parallel increasing nutrient content and temperature of the water and the decrease of the water flow and shallowing of the water body. to rank the suitability of the habitats in term of the mosquitoes we also used the data reported in tóth (2004). the typology of mosquito habitats is based on the size of the surface water, the depth of the riverbed and the connecting branches, and the abundance of the macro-vegetation. the correspondence between the topology systems by amoros et al. (1987) and tóth (2004) was based on the hydrological similarities of the water bodies (tab. 3). eupotamon a, b and the parapotamon a and b waters are relatively cold, deep, nutrient-poor and flowing envitab. 1. number of the sites where the mosquito larvae were collected in hungary according to the data reported by tóth (2004). species fp mr sl as os ls lp aw mn tn ta an. algeriensis 4 an. atroparvus 2 1 2 13 10 72 7 an. claviger 2 3 17 4 175 53 297 163 24 an. hyrcanus 9 3 19 2 1 8 23 3 7 an. maculipennis 9 20 21 16 318 28 550 267 33 an. messeae 6 4 1 3 1 29 13 110 24 13 ae. cinereus 4 6 14 4 164 33 269 192 11 ae. rossicus 1 17 1 26 5 ae. vexans 4 16 19 10 245 54 422 221 31 cx. modestus 7 4 11 31 14 330 69 312 46 10 cx. pipienspipiens 11 11 6 32 1 304 85 594 387 54 fp, floodplain and floodplain pools; mr, medium-sized rivers; sl, small lake reservoirs; as, artificial small lakes with standing waters; os, other types of small artificial lakes with standing waters; ls, littoral shallow lakes; lp, litoprofundal shallow lakes; aw, artificial swamps and standing waters; mn, marsh-type natural standing waters; tn, tömpöly-type; ta, natural waters, tömpöly-type artificial waters. tab. 2. number of the collected larvae (corresponding to w) in the different habitats (tab. 3) according to tóth (2004; tab. 1). habitat type/species an. alg. an. atr. an. cla. an. hyr. an. mac. an. mes. eupotamona 0 2 2 9 9 6 eupotamonb 0 1 1 4.5 4.5 3 parapotamona 0 2 2 9 9 6 parapotamonb 0 1 1 4.5 4.5 3 plesiopotamon 0 5.33 49.75 6.25 93.75 7.6 paleopotamon 4 29.67 134.25 10.25 219.5 40 bypass channel 0 2 2 9 9 6 cascadepotamon 2 15.33 67.63 7.38 112 21.5 habitat type/species ae. cin. ae. ros. ae. vexans cx. mod. cx. pip. eupotamona 4 0 4 7 11 eupotamonb 2 0 4 3.5 5.5 eupotamona 4 0 4 7 11 parapotamonb 2 0 4 3.5 5.5 plesiopotamon 47 9 304 96.5 70.8 paleopotamon 126.25 10.67 182 109.25 280 bypass channel 4 0 4 7 11 cascadepotamon 64.13 5.33 93 56.38 142.75 no nco mm er cia l u se on ly 51 a. trájer et al. ronments, which are almost unsuitable for aedes, anopheles and culex species. this is clearly reflected in the countrywide data collected by tóth (2004). eupotamon and parapotamon water bodies correspond to the floodplain, floodplain pool and medium-sized river environments. in contrast, plesiopotamon and paleopotamon habitats provide a relatively warm, shallow, calm and nutrient-rich environment, which is essential for mosquitoes (tab. 3). the habitat preference of the mosquitoes according to the number of larvae collection (tóth, 2004) reflects the classical belief that the larvae of anopheles, aedes and culex prefer marshes, swamps and backwaters. many species of larvae were mainly found in plesiopotamon and paleopotamon waters (an. claviger, an. maculipennis, ae. rossicus, ae. cinereus, ae. vexans, cx. modestus and cx. pipiens pipiens). the larvae of anopheles algeriensis were found only in paleopotamon-like habitats. the larvae of an. atroparvus, an. messeae and, particularly of an. hyrcanus are known in a larger numbers of water types, including colder and faster flowing potamon types. the habitat suitability-weighted area (hswa) was calculated by the multiplication of the area of the water body within the grid and the number of occasions in which the larvae of the given mosquito species were collected by tóth (2004) from the corresponding water types. hswa=aw where a=area of the water body within the grid; w=the habitat-related number of larvae according to tóth (2004). the calculation of the evenness species evenness (e) quantifies how equal the community is numerically. normally, the basic data of the calculation process is the relative abundance of a species and the number of the species (s). in this study, the hswa value was instead used to approximate the relative abundance of the species within the population according to the shannon-equation. species evenness is computed as: e= h’ lns where h’=shannon-index (eq. 3); s=number of the species (it is fixed; n=11). h’ is calculated as: where s=number of the species (it is fixed; n=11) pi=the proportion of the individuals of a species i, which is approximated and replaced with the corresponding hswa value. results the change of the weighted areal suitability of different mosquitoes the hswa value is not equivalent to the abundance or the proportion of a mosquito species in the mosquito fauna, tab. 3. the correspondence between riverbed types and the aquatic habitats. habitat types trapping habitats of mosquitoes according to tóth (2004) eupotamona floodplain and floodplain pool bypass channel floodplain and floodplain pool eupotamonb medium-sized river parapotamona floodplain and floodplain pool parapotamonb floodplain and floodplain pool plesiopotamon small lake reservoir types of artificial small lakes of standing waters other types of small artificial lakes of standing waters littoral shallow lake litoprofundal shallow lake paleopotamon artificial swamps and standing waters marsh-type natural standing waters tömpöly-type natural waters tömpöly-type artificial waters cascadepotamon transition between eub and parb (mid water-level) transition between eua and eub (high water level) transition between eub and plesiopotamon (low water level) nulpotamon drylandin a holm after of before the formation of a water body no nco mm er cia l u se on ly 52impact of river bank regulation on mosquito vectors but it is an area-based index, which can approximate the abundance of a mosquito species according to the suitability of an area as habitat. the relatively warm, shallow and nutrient-rich riverbed types are preferred by the different mosquito species. however, as shown in fig. 2, there are notable differences between the habitat preferences of the species. the results of the mosquito-suitability modeling of the habitats reflected the consequences of the antropogenic riverbed transformations (figs. 3 and 4). however, the results varied species by species according to the differences in habitat preferences. the results of the modeled runs showed three local minima for the species ae. cinereus, ae. vexans, an. claviger, an. atroparvus and an. messeae in 1830, 1946 and 1986. the weighted areas of cx. pipiens pipiens showed two major local minima in 1946 and 1986. an. algeriensis and an. maculipennis showed two major local minima in 1946 and 1986, and a minor in 1830. local minima were also apparent for an. hyrcanus and cx. modestus in 1830, 1946 and 2004. a major local minimum was found for ae. rossicus in 1830. an. atroparvus and an. algeriensis showed the greatest hswa values in 1790, 1955 and 2013. similar local maxima were found for an. algeriensis, an.claviger, ae. cinereus, ae. rossicus, ae.vexans and cx. modestus in 1790, 1955 and 2013. ae. maculipennis and cx. pipiens showed a largely decreasing trend during the studied years. the fluctuation of the maxima of an. messeae and an. hyrcanus showed individual patterns (figs. 3 and 4). the modeled mean area-based suitability index proportion of the mosquito species showed the dominance of cx. pipiens pipiens (%hswa: 42.2) and an. maculipennis (%hswa: 26.7) in the mosquito fauna (fig. 5). a few species showed intermediate % mean hswa values in the studied area, namely ae. vexans (7.5), cx. modestus (6.6), ae. cinereus (5.3), an. claviger (5.2), an. messeae (2.4) and an. hyrcanus (2.1). the %hswa value of the most notable potential malaria-vector mosquito an. messeae ranged between 1.9 and 2.6, showing relatively modest variations (sd: 0.26; fig. 5). the changes in the evenness values showed a significant increasing trend after the river regulation until the 1950s (p=0.014) (fig. 6). before the river regulation (1790-1870) the mean evenness was 0.62, whereas after the river regulation (1946-2013) the mean value was 0.74. the construction of the hydropower plant did not caused significant changes in the evenness values. during nearly a 200-year period, the river streams underwent important modifications, showing changes in their attribution to specific stream types or even disappearing or forming in other zones of the studied area (farkas-iványi and trájer, 2015). since the time periods between dates of the maps were different, the number of stream transitions was calculated for identical long periods. the frequency of the standardized transitions per year calculated to the half periods between 1805 to 2008 fig. 2. mean relative number of habitats where the mosquito larvae were found. eua; eupotamon a; eub, eupotamon b; para, parapotamon a; parb, parapotamon b; ple, plesiopotamon; pal, paleopotamon. an. alg, anopheles algeriensis; an. atr., anopheles atroparvus; an. clav., anopheles claviger; an. hyr., anopheles hyrcanus; an. mac., anopheles maculipennis; an. mes., anopheles messeae; ae. cin., aedes cinereus; ae. ros., aedes rossicus; ae. vex., aedes vexans; cx. mod., culex modestus; cx. pip., culex pipiens pipiens. no nco mm er cia l u se on ly 53 a. trájer et al. showed a pattern that was comparable to that of evenness (fig. 6). higher evenness values were related to higher transition numbers (significance of the correlation, p=0.039). discussion before river regulations, the continuous and natural change of the river bank system provided excellent ecological habitats for a number of aquatic organisms. the natural river bank dynamic was characterized by repeated floods that gave rise to the mosaic of the different habitats in the upper flow of the danube in hungary. the river occasionally left its bed, creating meanders and a rich channel network (hohensinner et al., 2004). the slight fluctuation in the proportion of the different aquatic habitats indicates that a balance existed between the renewal, the formation and the terrestrialization of the aquatic habitats. this balance in the dynamics of the river system changed and declined abruptly after the river regulation in the 19th century. the process culminated in an almost complete devastation of the marshes and swamps following the political and economic decisions to drain the wetlands in hungary after the 2nd world war. the building of the gabčíkovo-nagymaros waterworks also caused the fig. 3. modeled habitat suitability-weighted area of the different mosquito species on absolute hswa value scale. fig. 4. modeled habitat suitability-weighted area of the different mosquito species on log(hswa) value scale. no nco mm er cia l u se on ly 54impact of river bank regulation on mosquito vectors fig. 5. the modeled %hswa values of the mosquito species. anopheles hyrcanus was not marked in the chart diagram due to its low values. no nco mm er cia l u se on ly 55 a. trájer et al. substantial modification of the river system completing the truncation of the danube river. the area of eupotamal water bodies of the szigetköz suffered a similar, significant decrease after the channelization as in case of the austrian marchland (hohensinner, 2004). according to a few preliminary measurements of temperature and oxygen carried out on 14 august 2014 (trájer et al., personal communication), the sequence from the eupotamon a, b to parapotamon a, b, plesiopotamon and paleopotamon can be ranked along an environmental gradient, namely from cool and oxigen rich, to warmer and oxygen poor river stream types. this successive order also coincides with the suitability of a stream type for mosquitoes, which is accordance with the findings of tóth (2004). our habitat-matching models showed that, after the first river regulation, the number and the area of mosquito habitats increased until the mid-20th century. it may be debatable whether the regulated potamon ecotype corresponds to the original definitions by tóth (2004). however, the present eupotamon a and b types continue to designate freshwaters with greater inflow, whereas plesiopotamon and paleopotamon are warmer, shallow and nutrient-rich water bodies with little or no flow. our results are in accordance with the findings of vaňhara (1991), who found that, after the extensive water management alterations on the southern moravian dyje and morava rivers, the number and the diversity of the mosquito species of e.g. culex and anopheles in permanent water bodies increased. nevertheless, the observations of vaňhara (1991) spanned a shorter period, which could not reflect the mediumand short-term changes of the regulated potamon system. according to our model results, it is probable that, after the initial increase of the number and area of the shallow and permanent water bodies, the secondary changes of the clenched floodplain between the barriers and the deepening eupotamon a and b decreased the habitats suitable for the mosquito development. we have to note that our aim was to characterize the area-based mosquito larva suitability of a given area, not to estimate long-term changes in the number of mosquitoes. the hswa value cannot approximate the abundance of a mosquito species because it is based on larval collection. this method can therefore only approximate the relative number of the larvae in water bodies. though hswa can be considered a rough estimation of the potential relative number of adult mosquitoes, the predatory pressure, meteorological conditions and flood events can substantially change the number of imagoes in a given period. since mosquitoes prefer the plesiopotamon-paleopotamon habitat types, the change of the weighted suitable habitats of the species inversely correspond with the area of the eupotamon a, b and parapotamon a, b riverbeds. historically, since the austrian regulation allowed filling the major part of the branches in the danube system, the area of the suitable habitats began to decrease in the early 1800’s. this caused the decline of the habitats of an. atroparvus, an. algeriensis, an. maculipennis, an. claviger, ae. cinereus, ae. rossicus, ae. vexans and cx. modestus, whereas the habitats of an. hyrcanus and an. messeae slightly increased. the large-scale regulation of the river, fig. 6. evenness values of the mosquito species and the frequency of the standardized transitions per year (explanations in the text). no nco mm er cia l u se on ly 56impact of river bank regulation on mosquito vectors in 1870, caused an increase in the area of the paleopotamon and plesiopotamon waters. in turn, according to our model, the regulation increased the habitats of an. atroparvus, an. algeriensis, an. claviger, ae. cinereus, ae. rossicus, ae. vexans and cx. modestus. in the hydropower dam stage, the area of the mosquito suitable habitats increased due to the drastically decreasing groundwater level, which was advantageous particularly for an. atroparvus, an. algeriensis, an. messeae, an .claviger, ae. cinereus and ae. vexans. based on the modeled mean area-based suitability index, the relative abundance of the first four more frequent mosquito species showed the following order: cx. pipiens pipiens>an. maculipennis>ae. vexans>cx.modestus. this ranking is somewhat similar to the countrywide frequency order of the mosquito species in hungary, namely cx. pipiens pipiens> ae.vexans>>cx. modestus>an. maculipennis. the hswa value only characterize the mosquito suitability of the larger water habitats. however, since mosquitoes that develop in small water bodies like dendrotelmata or puddles were not considered, the hswavalues do not specify the real potential abundance. despite this limitations in the application of the method, the evenness based on hswa values correlated significantly with the transition numbers. this is an ecologically interpretable and important result since functional richness and diversity can reach their peaks at moderate disturbance frequency and intensity (biswas and mallik, 2010). while the calculated evenness with a fixed species number reached its peak decades after the river regulation, the construction of the hpp did not increase the evenness values. it is plausible that the relatively slow and moderate modifications in the fluvial network was less drastical than the construction of the hydropower plan, resulting a gradual change that increased the area and the diversity of water habitats. though the number of species was necessarilly fixed in this retrospective approach, other studies found a direct influence of disturbance on the evenness, without affecting the diversity of a studied group (kimbro and grosholz, 2006). in our study,the modifications of the river system triggered a chain of successional events that caused an increase in the evenness values and of the area more suitable for the development of the mosquito genera such as some member of anopheles. these changes may increase the risk of an intensification in the recurrence of malaria. ultimately, this risk may be also enhanced by the effects caused by climate change (martens et al., 1999). conclusions the quasi natural river dynamics may have a protective effect against the excessive habitat colonisation by mosquito species. our findings suggest that the anthropogenic river bank modification strongly influenced the composition of the successive stages even after 60-70 years after the start of the regulation. these changes caused an increase in the mosquito-suitable habitats, although the effect was positive or negative depending on breeding-habitat preference of the different mosquito species. according to the model, the increase of the evenness of the mosquito fauna increased after the first major disturbance (the primary river regulation), maintaining constant high values since the mid-20th century. the disruption of the natural stream system caused the increase of the hematophagous mosquito diversity. the increasing areal extension of relatively warm and nutrient-rich water habitats had also positive effects on the rarest members of the mosquito fauna, such as the potential malaria mosquito vectors. the renaturalization of the river system may provide a long-term tool for the prevention of mosquitoborne diseases increasing the relative and absolute area of the eupotamon and parapotamon riverbeds within the floodplain. this effect is somewhat controversial since the increasing riverbed dynamics can also maintain the permanent regeneration of the paleopotamon and plesiopotamon riverbeds. acknowledgements the study was supported by the támop 4.2.2.a11/1/konv-2012-0064: impact of extreme weather events on surface waters subproject. references ackerknecht eh, 1945. malaria in the upper mississippi valley 1760-1900. bull. med. libr. assoc. 33:372-373. amoros c, roux al, reygrobellet jl, bravard jp, pautou g, 1987. a method for applied ecological studies of fluvial hydrosystems. regul. river. 1:17-36. bartholy j, pongrácz r, gelybó gy, 2007. regional climate change expected in hungary for 2071-2100. appl. ecol. env. res. 5:1-17. bartholy j, pongrácz r, gelybó g, szabó p, 2008. analysis of expected climate change in the carpathian basin using the prudence results. quat. j. hung. met. serv. 112:249-264. benedict mq, levine rs, hawley wa, lounibos lp, 2007. spread of the tiger: global risk of invasion by the mosquito aedes albopictus. vector-borne zoonot. 7:76-85. biswas sr, mallik au, 2010. disturbance effects on species diversity and functional diversity in riparian and upland plant communities. ecology 91 28-35. brown h, diuk-wasser m, andreadis t, fish d, 2008. remotely-sensed vegetation indices identify mosquito clusters of west nile virus vectors in an urban landscape in the northeastern united states. vector-borne zoonot.8:197-206. cailly p, balenghien t, ezanno p, fontenille d, toty c, tran a, 2011. role of the repartition of wetland breeding sites on the spatial distribution of anopheles and culex, human disease vectors in southern france. parasit. vectors 6:65. no nco mm er cia l u se on ly 57 a. trájer et al. cajthaml j, 2013. polynomial georeferencing method for old map series. proc. 13th sgem geoconference on informatics, geoinformatics and remote sensing1:859-866. christensen n, wood aw, voisin n, lettenmaier dp, palmer rn, 2004. the effects of climate change on the hydrology and water resources of the colorado river basin. climatic change 62:337-363. dudgeon d, 1995. river regulation in southern china: ecological implications, conservation and environmental management. regul. river. 11:35-54. farkas-iványi, k, gábor g, 2013. historical changes of habitat dynamics in the szigetköz floodplain of the danube river. accessed on 20 september 2014. available from: http://is.muni.cz/repo/1093491/fbfw2013_final.pdf farkas-iványi k, guti g, 2014. the effect of hydromorphological changes on habitat composition of the szigetköz floodplain. acta. zool. bulgar. 66: (s.7):117-121. farkas-iványi, k, trájer, a, 2015. the influence of the river regulations on the aquatic habitats in river danube, at the bodak branch-system, hungary and slovakia. carpath. j. earth env. 10:235-245. fok é, 2007. the importance of dirofilariosis in carnivores and humans in hungary, past and present. mappe parassitologiche 8:181-188. fürst h, 2006. the hungarian-slovakian conflict over the gabčíkovo-nagymaros dams: an analysis.accessed on: august 2014. available from: http://www.columbia.edu/cu/ece/ research/intermarium/vol6no2/furst.pdf githeko, ak, lindsay, sw, confalonieri ue, patz ja, 2000. gcc and vector-borne diseases: a regional analysis. bull. world health organ. 78: 1136-1147. guti g, potyó i, gaebele t, weiperth a, 2010. ecological benchmarking of the aquatic habitat changes in the szigetköz floodplain of the danube. proceedings 38th iad conf., dresden, hungary. available from: http://www.academia.edu/ download/30341757/guti_et.al.szigetkoz_floodplain_38.ia d.201.pdf hohensinner s, habersack h, jungwirth m, zauner g, 2004. reconstruction of the characteristics of a natural alluvial river-floodplain system and hydromorphological changes following human modifications: the danube river (18121991). river res. appl.20:25-41. hohensinner s, haidvogl g, jungwirth m, muhar s, preis s, schmutz s, 2005. historical analysis of habitat turnover and age distributions as a reference for restoration of austrian danube floodplains. wit trans. ecol. envir. 3:489-502. hohensinner s, jungwirth m, muhar s, schmutz s, 2011. spatio-temporal habitat dynamics in a changing danube river landscape 1812-2006. river res. appl. 27:939-955. jetten th,takken w, 1994. anophelism without malaria in europe. a review of the ecology and distribution of the genus anopheles in europe. wageningen agricultural university, wageningen: 69 pp. kimbro dl, grosholz ed, 2006. disturbance influences oyster community richness and evenness, but not diversity. ecology 87 2378-2388. klobučar a, merdic e, benic n, baklaic žl, krčmar sa, 2006. first record of aedes albopictus in croatia. j. am. mosquito contr. 22:147-148. kovats rs, campbell-lendrum dh, mc michel aj, woodward ah, cox js, 2001. early effects of gcc: do they include changes in vector-borne disease? philos. t. r. soc. b 356:1057-1068. krisztalovics k, ferenczi e, molnar zs, csohan a, ban e, zoldi v , kaszas k 2008. west nile virus infections in hungary, august-september 2008. euro surveill.13:19030. lőrincz f, 1937. malaria in hungary. riv. malariol. 16:465-479. martens p, kovats rs, nijhof s, de vries p, livermore mtj, bradley dj, cox j, mc michael aj, 1999. climate change and future populations at risk of malaria. global environ. chang. 9:s89-s107. mushinzimana e, munga s, minakawa n, li l, feng cc, bian l, kitron u, schmidt c, beck l, zhou g, githeko ak, yan g, 2006. landscape determinants and remote sensing of anopheline mosquito larval habitats in the western kenya highlands. malaria j.5:13. nilsson c, reidy ca, dynesius m, revenga c, 2005. fragmentation and flow regulation of the world’s large river systems. science 308:405-408. overgaard hj, ekbom b, suwonkerd w, takagi m, 2003. effect of landscape structure on anopheline mosquito density and diversity in northern thailand: implications for malaria transmission and control. landscape ecol. 18:605-619. scholte ej, schaffner f, 2007. waiting for the tiger: establishment and spread of the aedes albopictus mosquito in europe, p. 241-260. in: w. takken and b.g.j. knols (eds.), emerging pests and vector-borne diseases in europe. wageningen academic publishers. szénási z, vass a, melles m, kucsera i, danka j, csohán a, krisztalovics k, 2003. malaria in hungary: origin, current state and principles of prevention. orvosi hetilap 144:10111018. tóth s, 2004. [magyarország csípôszúnyog-faunája (diptera: culicidae).6].[book in hungarian]. somogy megyei múzeumok igazgatósága, kaposvár: 327 pp. tőry k, 1952.[a duna és szabályozása].[book in hungarian] akadémiai kiadó, budapest: 454 pp. trájer aj, bede-fazekas á, bobvos j,páldy a, 2014. seasonality and geographical occurrence of west nile fever and distribution of asian tiger mosquito. quat. j. hung. met. serv. 118:19-40. vaňhara j, 1991. a floodplain forest mosquito community after manmade moisture changes (culicidae, diptera). regul. river 6:341-348. zittra c, kocziha z, pinnyei s, harl j, kieser k, laciny a, eigner b, silbermayr k, duscher gg, fok é, fuehrer hp, 2015. screening blood-fed mosquitoes for the diagnosis of filarioid helminths and avian malaria. parasite. vectors 8:1-6. no nco mm er cia l u se on ly layout 1 introduction the italian association of oceanography and limnology (aiol) invited me to summarize my work in the field of marine biology and ecology, a path that started in 1976 with a joint paper with norberto della croce (della croce and boero, 1976), one of the founders of aiol. necessarily, the number of self-citations will prevail in the references. i have tackled issues regarding biodiversity and ecosystem functioning (bef) in a series of papers that i will assemble here, synthesizing my own views on how to deal with bef. in most cases, i simply put together pieces of knowledge that were produced by other authors, that are cited in my papers and that i will not cite here, if not directly relevant. i will cite papers that i overlooked, while expressing as original what was not original at all. such as the papers that i did not cite in my contribution on fluctuations and variations in the marine environment (boero, 1994), namely fauvel (1901) and parenzan (1934). i like very much the say “you are what you eat”, and i extend it somehow to “you are who you meet”. i met many interesting people, and learnt a lot from them, absorbing their wisdom like a sponge. i spent long periods in places that are at the opposite ends of what we call civilization: california and papua new guinea. the places where we live are also important in shaping us. having met so many people, in so many countries, contributed to the way i see things. i want to cite a single person here: frank zappa. frank was a composer, and he formalized his approach to music with two concepts: the big note, and conceptual continuity. when we hear a piece of music played by an orchestra, we hear a big note. the “little notes” of the players, once blended, produce the music, perceived as a single combination of vibrations. if we listen to a single instrument at a time, we do not hear the music. the “little notes” are linked by a conceptual continuity that leads to what we hear. the composer writes the scores for each instrument so as to produce the big note as a final result. i like music and when i learnt about the way zappa saw it, i realized that science is the same. with reductionism, scientists split a complex reality into a suite of less complex objects that are investigated in isolation from the rest. this led to prodigious scientific advances in all fields of science, whose branches are just like the instruments of an orchestra, each one playing its own “little notes”. of course, each player thinks that s/he is playing the most important part of the music. after almost a century of specialization, we realized that the whole is more than the sum of the parts: the appreciation of emerging properties is the trademark of ecology. reductionism, then, should evolve into a holistic approach, identifying the conceptual continuity that links the little notes of each science branch into the big note, i.e., the natural world. there is a difference between music and science, though. the composer is the creator of the music. scientists analyse phenomena that are not the result of their action. steve vai is a guitar player that, listening to the intricate guitar solos improvised by frank zappa, transcribed them and produced scores that did not exist before. we are like steve vai, let us say. but we do not transcribe the scores played by a single instrument. music can be composed on a single instrument, to be then orchestrated. zappa composed the black page as a drum solo, and then he wrote the scores for a whole band, so as to enhance its “statistical density”. article linking ecosystems, habitats, and biodiversity: from the grand picture to the tiny details, and back ferdinando boero1,2,3 1stazione zoologica anton dohrn, naples, italy; 2university of naples federico ii, italy; 3cnr-ias, genova, italy abstract natural sciences usually proceed through the analysis of facts that are then assembled into a general framework, often called a “theory”. i have tried here to assemble the “tiny facts” that i have uncovered in my career and to organize them into a holistic perspective. i have chosen to start from the “big picture”, i.e., the functioning of ecosystems, to focus then on details regarding the expression of biodiversity, from the role of life cycles in ecosystem functioning, to the way of assessing biodiversity based on the accurate knowledge of its evolution in time. the historical biodiversity index allows to compare the potential biodiversity (all the species recorded from the studied habitat type) with the realized biodiversity (the species found by sampling in that habitat). the study of natural history might lead to unexpected ecological connections, such as the dynamics of plankton (the most important ecological phenomenon of the whole planet) and the composition of resting stage banks, or the keystone role of the interstitial fauna in determining the diversity of plankton. the oceanic realm is in three dimensions and must be considered as a volume rather than as an area. living systems, though, change constantly and a fourth dimension (time) is crucial to understand their structure and function. the cells of ecosystem functioning, based on connectivity, are proposed as natural spatial units for both management and protection from human impacts. no nco mm er cia l u se on ly linking ecosystems, habitats, and biodiversity: from the grand picture to the tiny details, and back 47 the challenge of this paper is to build a conceptual continuity that links all the things i have done in almost half a century. the problem, with science, is that we cannot listen to the message of 20 scientists, as we do when musicians play orchestral music. we listen to them one at a time. words cannot convey a holistic vision because they are necessarily reductionistic: we hear one word at a time. but figures can deliver a visual message that can be holistic. i drew figures in my papers but, when i met the artist alberto gennari, i stopped producing drawings and i asked him to transform my concepts into visual objects that might deliver a holistic message. how to depict an ecosystem the ocean covers 71% of the planet, but the oceanic space is not a surface, it is a volume. with an average depth of about 4.000 m, the volume of the ocean is more than 90% of the life-inhabited space. hence, the ocean is the rule, and the land is an exception, even though we tend to use the term ecosystem to label terrestrial ecosystems, and we use the adjective “marine” when dealing with oceanic ecosystems (boero, 2021a). if we want to depict the functioning of an ecosystem, then, it is scientifically sound to describe a marine ecosystem, considering mostly the water column, i.e., the threedimensional space that makes up the bulk of “terrestrial” ecosystems. the main characters that play a role in the functioning of an ecosystem and the relationships that link them to one another are illustrated in figure 1. visually, carnivores dominate the water column, as exemplified by the white shark eating the tuna, eating the mackerel, eating the sardine. such an environment cannot persist, even though fisheries science often focuses on fish and disregards other components of the ecosystems, for instance, jellyfish (boero, 2013). in the euphotic zone, the coastal landscape is dominated by algae and seagrasses, figure 1. the structure and function of an ecosystem (concepts: f. boero; artwork: alberto gennari; from the exhibit of the darwin dohrn museum, naples, italy). no nco mm er cia l u se on ly f. boero48 and by zoobenthos (not shown), but the vast majority of the water column, i.e., the pelagic domain, is not characterized by primary producers that we can see. the adult bony fish we see, apparently making up most of the whole trophic network, start their life as eggs, embryos, larvae, and juveniles. these early stages feed on herbivorous zooplankton (here exemplified by a copepod) that, in their turn, feed on phytoplankton (here shown as dinoflagellates and diatoms). these are the herbivores and the primary producers: contrary to terrestrial ecosystems, they are microscopic. jellyfish eat both the herbivores and the eggs, larvae, and juveniles of fish. if one tuna is considered as a life cycle, it is highly probable that jellyfish kill more tuna (when they are eggs, larvae and juveniles) than white sharks do. all living beings eventually die and are decomposed by bacteria that make nutrients available to phytoplankton. viruses are a cause of bacterial mortality. the “microbes” (i.e., phytoplankton, bacteria, and viruses) are the core of ecosystem functioning and are often labelled as the microbial loop. microbes, however, are not only a loop, but they are also the pillars of all ecosystems. the lower part of figure 1 shows the deep sea, where light is either absent or not sufficient for photosynthesis. the living matter produced “above” changes status and, after the death of the organisms, sinks towards the deep and is decomposed by bacteria, becoming particulate organic matter (marine snow). also, faecal pellets (not shown) are important in this transfer from the surface to the deep, as are the migrations of organisms that move towards the surface to search for food. marine snow sustains a host of detritivores and suspension feeders, here exemplified with pelagic crustaceans and benthic echinoderms and annelids. the large carcasses, in this case, killer whales, fall to the sea bottom and are consumed by scavengers. this dark domain is the vast majority of the space available for life. the oxygen is produced by the primary producers and is also dissolved to the water through airwater exchanges. both phenomena occur near the surface. downwelling currents bring the oxygen to the deep sea and trigger upwelling currents that bring nutrients from the deep to the surface, where they will sustain the primary producers. the scenario depicted in figure 1 can be further defined in terms of pathways (figure 2). the center of the figure represents the microbial loop: all living beings die (the central black circle with corpses) and are decomposed by heterotrophic bacteria that, in their turn, can be killed by viruses. bacterial decomposition leads to the production of nutrients that are used by phytoplankton (here as diatoms and dinoflagellates). nutrients are brought to the sea also by terrestrial runoffs. heterotrophic microbes (here as a ciliate protozoan) feed on the other microbes. for millions of years, life functioned in this way. it still can happen that microbes prevail, and the microbial loop leads to a microbial pathway, as is the case of red tides, i.e., abnormal proliferations of phytoplankters such as dinoflagellates (top right). the evolution of metazoa led to the consumption of microbes by filter feeders, such as the small crustacean, a copepod, that is at the margin of the microbial core. a second pathway (bottom right), thus, involves microbes, herbivores, fish larvae and juveniles that, in their turn, become adult fish and feed on each other. this is our favorite pathway and leads directly to us, through fisheries. a third pathway (bottom left) involves gelatinous herbivores, here depicted as a salp chain. when these animals develop huge populations, they overexploit the microbial component and compete with the copepods, thus having an impact on the pathway that leads to the fish. a fourth pathway (top left) involves gelatinous carnivores, here depicted as a cnidaria jellyfish and two ctenophores. these predators feed upon the herbivores and the eggs, larvae, and juveniles of fish, being themselves top predators. gelatinous plankton can be represented by huge populations that can re-direct the functioning of ecosystems, as argued by boero et al. (2008) and boero (2013). not all living matter is recycled and portions of it can become incorporated in the sediments, where carbon sequesfigure 2. the four pathways of living matter in marine ecosystems (concepts: f. boero; artwork: alberto gennari). no nco mm er cia l u se on ly linking ecosystems, habitats, and biodiversity: from the grand picture to the tiny details, and back 49 tration occurs (bottom center). figure 2 shows the pathways that characterize the photic zone. a fifth pathway is shown in figure 1 for the deep sea, where marine snow triggers a trophic network based on detritus, with species that can perform diel migrations from the deep sea to the surface and back. intrainterand extra-specific fluxes figures 1 and 2 can be formalized in an abstract fashion (without showing the organisms) so as to illustrate the fluxes of matter across food webs. ecology is usually divided into several branches that tackle single topics. life cycles, for instance, deal with the flux of living matter across generations of the same species (intraspecific fluxes). trophic networks, instead, deal with the flux of living matter from one species (the producer) to another species (the consumer) (interspecific fluxes) (boero and bonsdorff, 2007). figure 3 assembles three approaches to the study of the fluxes of matter across ecosystems. biogeochemical cycles represent fluxes of matter that is not organized in a living form, i.e., into species, and can be defined “extraspecific” fluxes. primary producers, plants, protists and monerans, with photoand chemosynthesis, organize the “nutrients” in a living form. these organisms, like all other organisms, perpetuate the existence of their species through life cycles that represent “intra-specific” fluxes across generations. the fate of primary producers is double. they die and are recycled by decomposers, represented by fungi and monerans that disassemble them, contributing to the production of “nutrients”, fuelling extraspecific fluxes. besides decomposing the materials, these organisms consume them also to synthesize their bodies (bacterial and fungal synthesis). all other components of food webs are subjected to the same fate if they are not consumed by organisms at higher trophic levels. decomposers, like all other components of the food webs, perpetuate themselves with intraspecific fluxes, i.e., their life cycles. the primary producers can also be consumed by secondary producers, i.e., animals and heterotrophic protists (protozoans). in this case, living matter flows from the primary to the secondary producers giving rise to “inter-specific” fluxes. the guts of secondary producers decompose the primary producers, often with the aid of heterotrophic bacteria, and then synthesize their food into their own bodies (zoo-synthesis) besides using their food as a source of energy for their body functions. secondary producers, i.e., the herbivores, are in their turn consumed by tertiary producers, the carnivores, and these can be consumed by higher trophic levels that can be again predators or, in alternative, parasites, i.e., micro-predators that do not kill their prey. in marine systems, the food webs can be very long, as shown in figure 1, where a white shark eats a tuna that eats a mackerel, that eats a sardine, and where the larvae of all bony fish eat herbivorous zooplankton that feeds upon phytoplankton. the same species, furthermore, can occupy much different trophic positions during its life cycle. a tuna, for instance, can be preyed upon by a white shark as an adult large fish, but, as a larva, it can be eaten by a jellyfish (figure 1 and 2). the diet of a tuna, moreover, can be based on herbivores at larval and juvenile stages, whereas it shifts to much higher trophic levels as size increases. the linkages between the three branches of ecology that are often kept separate, i.e., life cycles, trophic networks, biogeochemical cycles are illustrated in figure 3. focusing on important processes the functioning of oceanic ecosystems is the most important ecological process on earth, since the ocean covers more than 70% of the planet. contrary to terrestrial systems, marine systems function in pulses, with seasonal plankton blooms. figure 4 shows the seasonal pulses at the base of the functioning of marine systems. in spring, phytoand zooplankton proliferate (in figure 1 it is shown how “microbes” sustain higher trophic levels, dominated by carnivores). in summer, the plankton that has not been consumed by planktivorous organisms dies off and falls to the bottom. in fall, mild mixing triggers secondary plankton blooms. in winter, turbulence is maximal, and the nutrients are resuspended, so fuelling the spring blooms of phytoplankton that, in turn, fuel the zooplankton in the following spring. a series of papers (boero, 1994; belmonte et al., 1995; boero et al., 1996; marcus and boero, 1998, among others) complements the biogeo-figure 3. the fluxes of matter across ecosystems. no nco mm er cia l u se on ly f. boero50 chemical explanation of ecosystem functioning with a life cycle approach that recognizes a crucial role to benthic stages of planktonic organisms. many coastal plankters, in fact, are represented by huge biomasses for short periods, to disappear shortly thereafter. the traditional answer to the question “where are they when they are not there?” is “somewhere else” or, in alternative, “they are very rare and then become abundant again”. the fate of plankters that form true swarms and then disappear is not a mystery if jellyfish are concerned. boero et al. (2008) answered this question for gelatinous plankton and especially jellyfish. some jellyfish are holoplanktonic, but most of them have benthopelagic life cycles that involve both planktonic (jellyfish) and benthic (polyps) stages. these, however, were often studied by different scientists that focused either on the benthic or the planktonic stages. bouillon et al. (2006) tried to put some order in the hydrozoa, proposing a single classification that unites both polyps and medusae but, for many planktonic taxa, from copepods to rotifers, diatoms and dinoflagellates, benthic stages are given almost no ecological relevance. figure 4 also shows the benthic resting stages and introduces life cycles so as to complement the biogeochemical approach to the explanation of plankton dynamics. phytoplankton does not come from nutrients, and zooplankton does not come from phytoplankton! as illustrated in figure 3, biogeochemical cycles, life cycles and trophic networks must be considered altogether, so as to fully understand the functioning of ecosystems. do plankton, nekton and benthos really exist? my personal answer is no. these domains have been recognized based on the tools we use to study marine biology and ecology. jellyfish are collected in the water column, with plankton nets or plastic bags, whereas their polyps are collected from the sea bottom, with scuba diving or with dredges and grabs, and those who study the very same animals consider just one stage of the life cycle. the same is true for fish, studied by fisheries biologists as nektonic adults, whereas their larvae are studied by planktonologists. benthic organisms with planktonic larvae are almost the rule. even if the larval and juvenile stages are often short-lived, they make choices that are vital for the benthic adults that develop from them (fraschetti et al., 2003). pati et al. (1999), furthermore, demonstrated the coexistence of the meiofauna and the resting stages of plankters in marine sediments. meiofauna and resting stages, however, are extracted with different techniques by different specialists, and are rarely considered as a whole. consequently, the relationships that might occur among these coexisting species are almost entirely unknown. many members of the meiofauna do have piercing mouthparts and sucking pharynxes: what do they pierce and suck? resting stages are “seeds” and, at the same time, are rich in nutrients and, thus, represent a potential food source for the meiofauna. pati et al. (1999) hypothesized that, as shown in figure 5, after a plankton bloom the resting stages of the dominating species would fall on the sea bottom and become incorporated into the sediments. the resulting massive amounts of resting stages might guarantee the re-occurrence of the bloom in the following favourable seasons, thus acting as “seeds”. this is seldom the case, and the meiofauna might play a keystone role with predation on the resting stages (which thus are also a food source) of the previously dominant species, so enhancing or the diversity of plankton. the links among plankton, benthos and nekton are pervasive, but the connections among these domains are figure 4. plankton seasonality. a: spring; b: summer; c: fall; d: winter (concepts: f. boero; artwork: alberto gennari). figure 5. the predation of meiofauna on resting stage banks (concepts f. boero; artwork alberto gennari). no nco mm er cia l u se on ly linking ecosystems, habitats, and biodiversity: from the grand picture to the tiny details, and back 51 often neglected and this artificial separation hinders a holistic comprehension of the functioning of marine ecosystems. boero (e.g., 1994, 1996, 2021a, 2021b) and boero et al. (2004, 2019b) have repeatedly stressed the need for the unification of these approaches which, however, remain separated from each other in most research projects. from ecosystem functioning to biodiversity and back the former section depicts the functioning of ecosystems in terms of connections through “artificial” compartments, defined for ease of analysis and seldom united into a single, holistic vision. this functional approach must be complemented with a structural counterpart. in recent decades the acronym bef (biodiversity and ecosystem functioning) became fashionable, with the aim of linking structure and function (heip et al., 2009). this is realized in the marine strategy framework directive of the eu and the eleven descriptors of good environmental status (ges) therein. the pillars of ges are just biodiversity and ecosystem functioning (boero, 2016). the holistic approach defined above is the result of reductionistic analyses, but i preferred to sketch the whole picture before entering structural details, even if these are the factual basis for the synthesis. biodiversity the assemblage of species living in a given ecosystem, often colonizing different habitats throughout their life cycles, is the most obvious definition of biodiversity. each species has an inner diversity in terms of genetic make ups, and the different combinations and abundances of species produce a higher level of diversity, with emerging properties that do not equal to the simple sum of the species that can be found in a given ecological space. this has led to many diversity indexes, translating into numbers the information that can be drawn by sampling at a given place. biodiversity can be considered as a “structure” that is conducive to a “function”, the functioning of ecosystems the previous section dealt with. the acronym bef (biodiversity and ecosystem functioning) links structure and function, patterns, and processes. since the times of darwin, it is assumed that high biodiversity leads to high efficiency in the functions of ecological systems (boero, 2015a). the aim of a bef approach is to reach a holistic vision of the marine environment, assembling reductionistic approaches so as to understand the whole after having carefully inspected the parts. this stems from the say that “the whole is more than the sum of the parts”, namely the emerging properties that should be the main object of ecological studies. at present, however, marine ecology is still mostly reductionistic. reductionistic ecology: the oxymoron marine science is still practised by many sciences. each branch focuses on portions of the marine space and of marine biodiversity and is often disconnected from other branches. the array of topics is overwhelmingly vast. physical oceanography deals with the physics of the sea. it is studied with oceanographic vessels, automated vehicles, and satellites and it generates predictive models. the interaction with the atmosphere is often disregarded, as is the interaction with the sea bottom and the coastline. this calls for marine geology, whose main objective is to map the sea bottom and understand sediment transport along coasts. this involves physics, but it is often the case that geologists collaborate with hydraulic engineers, rather than with physical oceanographers. marine chemistry deals with the chemical properties of seawater, firstly salinity which, indeed, interacts with temperature in generating thermohaline currents, such as the great ocean conveyor belt. ocean acidification is also a matter of chemistry, and is an emerging field of marine chemistry, with a great bearing on marine biology. marine biology, furthermore, is divided into three main domains: plankton, benthos, and nekton. they are further divided into many sub disciplines. phytoplankton comprises autotrophic organisms, such as diatoms, flagellates, and bacteria: these tiny organisms are called microbes, whose study is labelled as microbial ecology. zooplankton is divided into size categories that span from heterotrophic ciliates to the giants of macro-zooplankton which, in their turn, are also labelled as gelatinous plankton, comprising species that range from chordates (e.g., the thaliacea) to cnidarians and ctenophores. the plankton, furthermore, comprises also species that spend just a period of their life in this domain, sometimes as larvae (e.g., fish and many benthic animals) sometimes as adults (e.g., the sexually competent jellyfish that have benthic polyps). the nekton is made of animals that can actively swim against the current. the nekton is also studied by fisheries biology, from a resource management point of view. marine conservation deals with the protection of the marine environment. all these disciplines are based on the recognition of species, i.e., on taxonomy, whose approaches can be either phenotypic or genotypic, or both. the study of species must reconstruct their life cycles, and this clearly show that these domains are fading into each other for most species. many supposed holoplankters (spending their whole life in the plankton) very often do have benthic resting stages and their blooms are fuelled no nco mm er cia l u se on ly f. boero52 by a benthic resting stage bank (boero et al., 1996; marcus and boero, 1998). trophic networks, in their turn, connect the three domains. juvenile fish do feed on plankton, then change their diet when they grow up. as larvae or juveniles, fish are eaten by predators that then can be the prey of the adult stage. furthermore, the division between the deep sea, the high seas, and the coastal areas, each studied by different portions of the scientific community, is unnatural, since these domains are highly connected by vertical and horizontal currents, from the great ocean conveyor belt to the local upand down-welling processes generated by winds and submarine canyons. the ocean is, instead, one. such a need for a holistic approach is often invoked, as the european marine board did in the introduction of navigating the future iv: “to truly progress this knowledge, european scientists across a broad range of disciplines and domains must make a quantum leap towards holistic approaches and integrated research on a scale which will help us to much better understand, protect, manage and sustainably exploit the seas and oceans which surround us. this is a grand challenge; not just for europe, but for human society as a whole” (arnaud et al., 2013). the evolution of approaches the habitat approach embraced with the habitats directive, dealing exclusively with benthic habitats (fraschetti et al., 2008), eventually evolved into the ecosystem approach in the marine strategy framework directive and its 11 descriptors of good environmental status, having bef as its pillar (boero, 2016). “biodiversity is maintained” is the first descriptor. the other 10 list a series of impacts that must be kept below threshold, so as to maintain the functioning of the ecosystems in “good” status. of course, the meaning of “good” in these measures is: good for us. so, the concept of goods and services emerged with ecological economics: nature is important because it supports us with goods and services that can be given a price. if we destroy nature, we will lose money! the argument here is that decision makers care only for money, so it is better to talk about money when talking about nature, often confusing value with price (boero, 2008). the words “maintain” and “conservation” imply that we should identify a reference status and that we should keep it. but what is the reference status we want to maintain? the present one? almost all agree that the present one is not so good. so, the question is: what is the ocean we want? (boero, 2021b). do we want to maintain the current situation, or do we want it to go back to previous statuses? is it wise to propose that things must not change, or that they must go back to previous statuses? this is what political conservation wants. the natural world, however, is a changing world, whose trademark is evolution, i.e., change. a challenge, thus, is to understand change, and to study it. the marine realm has three dimensions, and must be studied in terms of volumes, but the processes studied by marine ecology are rapid. the water column makes up most of the ecological space available to life and living processes proceed in pulses: phytoplankton pulses are followed by zooplankton pulses, and the two sustain an apparently stable nektonic component. microbial processes are the core of the functioning of marine ecosystems and sustain most of the deep-sea life, where photosynthesis is impossible due to lack of light. marine ecosystems, thus, are very dynamic. as terrestrial animals, we are attracted by landscapes that are part of our experience, such as algal, seagrass or animal forests that are as stable as most terrestrial vegetation, but the majority of the marine space is not linked to the sea bottom and there the processes are very dynamic. fluctuations and variations: the fourth dimension the perception that natural systems do change within the short term, especially in the marine realm, is not new, and fauvel (1901) and parenzan (1934) were among the first to realize that even apparently stable benthic systems do vary in their species composition and that species that were previously rare can become suddenly abundant, whereas abundant species can become rare. boero (1994, 1996) reached their same conclusion, arguing that stability does not exist in natural systems, especially in marine ones, and that the situations we can depict with a given sampling session represents a “moment” that is part of a “story”. regular fluctuations (e.g., seasonal ones) are different from variations, when what is perceived as “normal”, in terms of the species composition at a specific location, changes and common species become rare whereas rare species become abundant. boero (1994) argued that the species pool in a given basin tends to remain stable, but that each species contributes with different biomasses to the total biomass that the system can “express” which can remain stable in terms of quantity but not of quality. furthermore, the species pool can change due to the arrival of non-indigenous species that can contribute to the maintenance of biodiversity (in terms of species numbers) in a changing ocean (boero, 2021b). if a basin warms up, due to global warming, the indigenous species will suffer from the new conditions, but these will be conducive to the thriving of species that are adapted to higher temperatures. these species are often non-indigenous ones. this calls for the second descriptor of good envino nco mm er cia l u se on ly linking ecosystems, habitats, and biodiversity: from the grand picture to the tiny details, and back 53 ronmental status: “non-indigenous species do not adversely alter the ecosystem”. non-indigenous species are often labeled as noxious invaders, but they can also replace the species that cannot withstand new conditions, determined by global change, as boero (2021b) suggested. non-indigenous species, or aliens, can be noxious, such as the ship driven ones, that invade basins where they can outcompete indigenous species (boero, 2002), but this is not the case for all “aliens”. the alteration of ecosystems, thus, can be due to changing physical conditions (e.g., global warming) or to the arrival of a noxious alien, as is the case of the ctenophore mnemiopsis leydi that dramatically changed the black sea ecosystem (boero et al. 2008; boero, 2013). boero et al. (2019), hence, proposed the concept of the ocean in 4d: the surfaces of the sea bottom and sea surface are in 2d, the water column is in 3d, and the dynamics of marine systems add a fourth dimension to the system: time. boero et al. (2015) stressed the importance of time series and labeled time as an affliction because, when a given situation seems to be described and understood... it changes, and our predictions fail. the historical nature of bio-ecology is not conducive to predictions stemming from mathematical models, at the press of a button (boero et al. 2004): we cannot predict what will happen tomorrow even in our history which, in turn, depends on the history of natural systems, with innumerable feedbacks and interacting variables. indeed, episodic events (boero, 1996) can be important drivers of change that lead to different statuses of ecosystem structure and function. measuring biodiversity, linking it to habitats boero and bonsdorff (2007) proposed the historical biodiversity index (hbi) to introduce time in the evaluation of the status of biodiversity. its application requires a thorough knowledge of biodiversity from a historical perspective. the index is very simple: hbi = realized biodiversity / potential biodiversity where realized biodiversity is the species pool resulting from a sampling session in a given habitat, and the potential biodiversity is the list of species that, in the biogeographical region where the samples have been collected, includes all species recorded in previous studies of the sampled habitat. checklists of species have been made for most biogeographic regions, the european register of marine species (costello et al., 2001) being a significant example. some work is still necessary to assemble a complete list of habitats (fraschetti et al., 2008) since benthic habitats have received much attention, whereas pelagic ones are still less clearly defined. for each biogeographic unit, thus, a matrix can be built, comprising both the species that have been recorded from that unit, and the habitats that are present in the very same unit. each species is then to be assigned to the habitats where it has been recorded from. some are exclusive to a single habitat (e.g., the species leaving only on seagrass leaves, or in specific symbiotic associations) whereas others are present across many habitats. the table leads to a species list for each habitat: the potential biodiversity. a sampling session in a given habitat leads to a species list: the realized biodiversity. this can be increasingly efficient also by using environmental dna. the value of hbi is 1 if the number of species found in sampling session (realized biodiversity) equals the number of species found throughout the study of the sampled habitat (potential biodiversity). if no species is found, the index is 0. the more the index approaches to 1, the more biodiversity is maintained, the more it approaches to 0, the more biodiversity is eroded. since the quantitative contribution of species varies in time (boero, 1994) it is useless to focus much on the quantities of each species, since they undergo continuous change. the apparent disappearance of a previously dominant species might not be a tragedy if other species replace it. besides being assigned to one or more habitats, furthermore, each species should be classified based on its trophic role. the presence of species that have a basic role in trophic networks should be more constant, whereas the species occupying high trophic levels are more liable of being “eroded”. the disappearance of top predators, and the dominance of species with lower trophic roles, is now labelled as trophic downgrading (britten et al., 2014). the comparison of species lists, if accurate, provides a wealth of information about the history of a given ecosystem, and of its status. the index, furthermore, provides a very useful information that is usually not considered in sampling sessions: it provides the list of species that have not been found, and that were found in the past. this information will allow the detection of species that are not being found anymore and that might be putatively extinct species (boero et al., 2013; boero and gravili, 2013; gravili et al., 2015). if the sampling leads to species that were previously unrecorded (e.g., non-indigenous species) they are to be added to both the potential and the realized biodiversity. since the exploration of biodiversity is far from being complete, it is possible that species new to science are also found. these might be indigenous species that have been overlooked by previous sampling sessions, or non-indigenous species that have not been described from their sites of origin. the quite evident jellyfish pelagia benovici, for instance, recently bloomed in the adriatic no nco mm er cia l u se on ly f. boero54 sea and was described as a new species by piraino et al. (2014). since the jellyfish fauna of the adriatic sea is rather well known, it was hypothesized that p. benovici reached the adriatic from elsewhere, even though, from a zoological point of view, the adriatic is its typical locality. the revision of genera, furthermore, might lead to the splitting of a nominal species into a set of distinct species, as is the case of aurelia (scorrano et al., 2017), whose diversity in the mediterranean is much greater than previously recognized. the species pool of a basin is made of all the species that have been recorded from its waters. some species are known since the beginning of the study of biodiversity and are part of what is considered as the “native” species pool of a basin. many of these species have been described in 1758 and in the following years, by the pioneers of the study of biodiversity, and they are usually quite evident. the original species pool is not very rich, but it increases with the refinement of biodiversity exploration. as long as species are discovered either as new species or as new records, the species pool increases in size. the hbi can, thus, identify putative extinctions. maritime spatial planning once biodiversity and ecosystem functioning are framed into a conceptual framework, they must also be framed spatially, with the four-dimensional approach depicted above. this will allow to manage our activities in the marine environment, as required by the european commission with the directive on maritime spatial planning. the habitats directive is a prelude to maritime spatial planning, since it identifies the benthic habitats that deserve protection. single states, furthermore, defined other important areas as national marine protected areas, not necessarily based on the habitats directive. boero (2017) argued that most mpas are designed to protect charismatic habitats that do not necessarily comprise ecosystems. it is rarely the case, in fact, that a single habitat comprises an entire ecosystem. protecting a habitat without protecting the ecosystem it depends on is not ecologically sound. for this reason, the european commission shifted its focus from habitats and mpas to ecosystems (the marine framework directive and the maritime spatial planning directive) (boero et al., 2016). this approach led to the definition of “cells of ecosystem functioning” (boero et al., 2019), based on ecological connectivity (figure 6). the cells of ecosystem functioning (cefs), thus, are the natural units of management and conservation, and should be at the basis of maritime spatial planning. the greatest challenge for marine science, as the sum of the various marine sciences, is to join forces and map the marine environment from both a structural and a functional point of view, i.e., mapping the cells of ecosystem functioning. conclusions i have omitted here my direct contribution to the exploration of marine biodiversity, culminated with two monographs on the hydrozoa of the mediterranean and of the world (bouillon et al., 2004, 2006). the writing of such monographs is increasingly rarer, due to the unwise dismissal of traditional taxonomy (boero, 2001, 2010) with the claim that genotypic approaches might replace phenotypic ones (boero and bernardi, 2014): both are necessary. the sample of my scientific production cited here does not cover all the topics i tackled in my career. the apparent jumps from one topic to the other(s) occurred almost by chance, due to opportunities or, even, to the calls for projects that forced me to wrap up what i was doing, discovering a posteriori that what i did had emerging properties. i would be a liar if i would say that i knew since the beginning where i wanted to go with my research. i focused on a specific topic, the polyp stage of the hydrozoa, then i had to deal with the medusa stage, passing from benthos to plankton, to arrive to nekton while considering the impact of jellyfish predation on fish larvae. the study of marine protected areas led me to shift from the habitat to the ecosystem approach, and this led to the natural link of all facets of marine science. i jumped on “strange” phenomena, when they occurred, such as the mass mortalities of benthic organisms, linking them to climatic fluctuations (rivetti et al., 2014), as i did with the changes in the hydroid assemblage that i studied in my youth (boero and figure 6. a hypothetical cell of ecosystem functioning in which currents link different habitats through propagule transport (the water column is also a suite of habitats) (concepts: f. boero; artwork: alberto gennari). no nco mm er cia l u se on ly linking ecosystems, habitats, and biodiversity: from the grand picture to the tiny details, and back 55 fresi, 1986) and that was the re-analized after decades (puce et al., 2009). philosophy of science attracted me since the very beginning of my career, but it took me a long time to understand that the popperian paradigm of falsification, calling scientists to produce universal statements (that can be only falsified and cannot be verified) is not valid in historical disciplines (such as ecology and evolution) that are based on existential statements (that can be only verified and cannot be falsified) (boero et al., 2004). gradual evolution, for instance, falsifies the universality of punctuate evolution, but not its existence (and vice-versa). dealing with historical systems (dominated by existential statements) as if they were a-historical systems (dominated by universal statements) is hindering the development of both ecology and evolution. despite the impossibility of performing predictions about the behaviour of ecological systems, it is anyway possible to understand them and to depict future scenarios (boero, 2015b), and to act so as to drive future ecosystems into a desired status (boero, 2021b). corresponding author: ferdinando boero, stazione zoologica anton dohrn, villa comunale, 80121 napoli, italy. e-mail: ferdinando.boero@unina.it conflict of interest: the author declares no potential conflict of interest. funding: none. availability of data and materials: all data generated or analyzed during this study are included in this published article. key words: ecosystem, habitat, biodiversity, ocean. received: 12 december 2022. accepted: 15 december 2022. publisher’s note: all claims expressed in this article are solely those of the authors and do not necessarily represent those of their affiliated organizations, or those of the publisher, the editors and the reviewers. any product that may be evaluated in this article or claim that may be made by its manufacturer is not guaranteed or endorsed by the publisher. ©copyright: the author(s), 2022 licensee pagepress, italy advances in oceanography and limnology, 2022; 13:11079 doi: 10.4081/aiol.2022.11079 this work is licensed under a creative commons attribution noncommercial 4.0 international license (cc by-nc 4.0). references arnaud s, arvanitidis c, azollini r, austen m, balguerias e, boero f, et al., 2013. navigating the future iv. position paper 20 of the european marine board. ostend, belgium. 203 pp. belmonte g, castello p, piccinni mr, quarta s, rubino f, geraci s, boero f, 1995. resting stages off the italian coast. in a elefteriou, ad ansell, cj smith (eds). biology and ecology of shallow coastal waters. olsen & olsen, fredensborg, denmark. p. 53-58. boero f, 1994. fluctuations and variations in coastal marine environments. p.s.z.n.i: marine ecology 15:3-25. boero f, 1996. episodic events: their relevance in ecology and evolution. p.s.z.n.i: marine ecology 17:237-250. boero f, 2001. light after dark: the partnership for enhancing expertise in taxonomy. trends in ecology and evolution 16:266. boero f, 2002. ship-driven biological invasions in the mediterranean sea. in: alien marine organisms introduced by ships in the mediterranean and black seas. ciesm workshop monographs 20:87-91. boero f, 2008. price and value, alternatives to biodiversity conservation (in the seas). ciesm workshop monograph 37:61-64. boero f, 2010. the study of species in the era of biodiversity: a tale of stupidity. diversity 2:115-126. boero f, 2013. review of jellyfish blooms in the mediterranean and black sea. gfcm studies and reviews 92: 53 pp. boero f, 2015a. from darwin’s origin of species towards a theory of natural history. f1000 prime reports, 7:49. boero f, 2015b. the future of the mediterranean sea ecosystem: towards a different tomorrow. rendiconti lincei 26:3-12. boero f, 2016. marine biodiversity and ecosystem functioning: the pillars of good environmental status. biologia marina mediterranea 23:50-57. boero f, 2017. from marine protected areas to mpa networks. in: p goriup (ed). management of marine protected areas. a network perspective. wiley and sons, chichester, uk. p. 1-20. boero f, 2021a. mission possible: holistic approaches can heal marine wounds. adv. mar. biol. 88:19-38. boero f, 2021b. the future ocean we want. advances in marine biology 90:51-63. boero f, belmonte g, fanelli g, piraino s, rubino f, 1996. the continuity of living matter and the discontinuities of its constituents: do plankton and benthos really exist? trends in ecology and evolution 11:177-180. boero f, belmonte g, bussotti s, fanelli g, fraschetti s, giangrande a, et al., 2004. from biodiversity and ecosystem functioning to the roots of ecological complexity. ecological complexity 2:101-109. boero f, bernardi g, 2014. phenotypic vs genotypic approaches to biodiversity, from conflict to alliance. marine genomics 17:63-64. boero f, bonsdorff e, 2007. a conceptual framework for marine biodiversity and ecosystem functioning. marine ecologyan evolutionary perspective 28:134-145. boero f, bouillon j, gravili c, miglietta mp, parsons t, piraino s, 2008. gelatinous plankton: irregularities rule the world (sometimes). mar. ecol. progr. ser. 356:299-310. boero f, foglini, f, fraschetti s, goriup p, machpherson e, planes s, soukissian t, coconet consortium. 2016. coconet: towards coast to coast networks of marine protected areas (from the shore to the high and deep sea), coupled with sea-based wind energy potential. sci-res.it 6:1-95. no nco mm er cia l u se on ly f. boero56 boero f, de leo f, fraschetti s, ingrosso g, 2019. the cells of ecosystem functioning: towards a holistic vision of marine space. advances in marine biology 82:129-153. boero f, kraberg, ac, krause g, wiltshire kh, 2015. time is an affliction: why ecology cannot be as predictive ad physics and why it needs time series. journal of sea research 101:12-18. boero f, cummins v, gault j, huse g, philippart c, schneider r, et al., 2019b. navigating the future v: marine science for a sustainable future. position paper of the european marine board, ostend, belgium. 24:1-89. boero f, carlton j, briand f, kiessling w, chenuil a, voultsiadou e, et al., 2013. marine extinctions. patterns and processes. ciesm workshop monographs 45:5-19. boero f., di camillo c., gravili c. 2005. aquatic invasions: phantom aliens in mediterranean waters. marbef newsletter 3, 21–22. boero f, fresi e, 1986. zonation and evolution of a rocky bottom hydroid community. p.s.z.n.i: marine ecology 7:123150. boero f, gravili c, 2013. the bio-ecology of marine extinctions, with a lesson from the hydrozoa. ciesm workshop monographs 45:75-79. bouillon j, medel md, pagès f, gili jm, boero f, gravili c, 2004. fauna of the mediterranean hydrozoa. scientia marina 68:1-449. bouillon j, gravili c, pagès f, gili j-m, boero f, 2006. an introduction to hydrozoa. mémoires du musèum national d'histoire naturelle. paris, france: muséum national d'histoire naturelle. 598 pp. britten gl, dowd m, minto c, ferretti f, boero f, lotze hk, 2014. predator decline leads to decreased stability in a coastal fish community. ecology letters 17:1518-1525. costello mj, emblow c, white rj (eds), 2001. european register of marine species: a check-list of the marine species in europe and a bibliography of guides to their identification. collection patrimoines naturels, 50. muséum national d'histoire naturelle: paris, france. 463 pp. fauvel p, 1901. les variations de la faune marine. feuille des jeunes naturalistes 363:78-81. della croce n, boero f, 1976. ecologia e biologia dei porti del mar ligure e alto tirreno. aspetti termici del golfo di la spezia. iiies journées d’etudes pollutions: 125-131. fraschetti s, giangrande a, terlizzi a, boero f, 2003. preand post-settlement events in hardand soft-bottom community dynamics. oceanologica acta 25:285-296. fraschetti s, terlizzi a, boero f, 2008. how many habitats are there in the sea (and where)? journal of experimental marine biology and ecology 366:109-115. gravili c, bevilacqua s, terlizzi a, boero f, 2015. missing species among mediterranean non-siphonophoran hydrozoa. biodiversity and conservation 24:1329–1357. heip c, hummel h, van avesaath p, appeltans w, arvanitidis c, aspden r, et al., 2009. marine biodiversity and ecosystem functioning. printbase, dublin, ireland. 91 pp. marcus n, boero f, 1998. production and plankton community dynamics in coastal aquatic systems: the importance of benthic-pelagic coupling and the forgotten role of life cycles. limnology and oceanography 43:763-768. parenzan p, 1934. rotazione biologica naturale. bollettino di zoologia 5:137-144. pati ac, belmonte g, ceccherelli vu, boero f, 1999. the inactive temporary component: an unexplored fraction of meiobenthos. marine biology 134:419-427. piraino s, aglieri g, martell l, mazzoldi c, melli v, milisenda g., et al., 2014. pelagia benovici sp.nov. (cnidaria, scyphozoa): a new jellyfish in the mediterranean sea. zootaxa 3794:455-468. puce s, bavestrello g, di camillo cg, boero f, 2009. longterm changes in hydroid (cnidaria hydrozoa) assemblages: effect of mediterranean warming? marine ecology-an evolutionary perspective 30:313-326. rivetti i, fraschetti s, lionello p, zambianchi e, boero f, 2014. global warming and mass mortalities of benthic invertebrates in the mediterranean sea. plos one 9:e115655. scorrano s, aglieri g, boero f, dawson mn, piraino s, 2017. unmasking aurelia species in the mediterranean sea: an integrative morphometric and molecular approach. zoological journal of the linnean society 180:243-267. no nco mm er cia l u se on ly layout 1 introduction estuarine and coastal environments are often characterized by energy sources for consumers which are heterogeneous and characterized by with large spatial and temporal variability (stowasser et al., 2012). the trophic interactions are driven by a complex array of multiple biological, chemical and physical processes, which altogether make trophic linkages among different groups within the aquatic food web difficult to be defined (layman et al., 2012; stowasser et al., 2012). such a difficulty in tracking the flow of energy along aquatic trophic webs is due to the fact that many species grow more than five orders of magnitude than others and some pass through several trophic levels during the different stages of their life cycle (cushing, 1975; pope et al., 1994; post, 2002; van oevelen et al., 2012; middelburg, 2014). food web studies carried out previously relied on gut content analysis of higher trophic level organisms (hall and raffaelli, 1993), and this method has a limitation in assessing the assimilation of material in the gut (stowasser et al., 2012). stable isotopic composition of carbon (δ13c) and nitrogen (δ15n) can be used as complementary tools to evaluate the structure and dynamics of ecological communities (peterson and fry, 1987; france, 1995; vander-zanden et al., 1997; post, 2002; middelburg, 2014; hinz et al., 2017). these isotopes provide information about the source of material in the integrated temporal scale (post, 2002), as δ13c and δ15n in the tissues of predators is enriched relative to their prey, and, thus, can be used to estimate trophic levels within a certain trophic web (minagawa and wada, 1984; jennings et al., 2002). the δ13c of consumers are usually close to that of their diet (<0.5‰; post, 2002) whereas δ15n is enriched in the consumers relative to their diet by 3.4‰ (deniro and epstein, 1978, 1981; minagawa and wada, 1984; post, 2002). dissimilarity in isotopic composition of δ13c and δ15n of marine organisms are caused by selective uptake of nutrients, feeding (meili et al., 1996; matthews and mazumder, 2003), and metabolic activities (deniro and epstein, 1981; hobson and clark, 1992; hobson et al., 2002; matthews and mazumder, 2006; santer et al., 2006). hence, the sensitivity of trophic position estimation carried out using stable isotope signatures depends on δ15n and δ13c of end-members, which, in turn, depends by several factors such as variation in source of nutrients, dissolved inorganic carbon (dic), organic matter etc. obtaining an appropriate baseline is therefore one of the most challenging methodological issues facing the effective application of stable isotopes to trophic food advances in oceanography and limnology, 2018; 9(1): 1-12 article doi: 10.4081/aiol.2018.7266 this work is licensed under a creative commons attribution-noncommercial 4.0 international license (cc by-nc 4.0). influence of river discharge on zooplankton diet in the godavari estuary (bay of bengal, indian ocean) jayati mukherjee,1 sura a. naidu,1 v.v.s. s. sarma,1* tuhin ghosh2 1csir-national institute of oceanography, regional centre, 176 lawsons bay colony, visakhapatnam; 2jadavpur university, school of oceanographic studies, kolkata, india *corresponding author: sarmav@nio.org abstract in estuaries, detrital (i.e., non-living) organic matter (om) contributes significantly to the particulate organic matter (pom) pool and we hypothesize that it may be a major source of estuarine zooplankton diet. to test this hypothesis, the isotopic composition of carbon (δ13c) and nitrogen (δ15n) of phytoplankton, zooplankton, and pom was assessed in the godavari estuary (bay of bengal, indian ocean) during wet (november) and dry periods (january). as a result of higher riverine discharge, pom concentrations and values of the c/chl-a ratio during the wet period were higher than those measured during the dry one. relatively lower δ13cpom values were observed during wet than dry period and contrasting to that was found for δ15npom. detritus from fresh water algae and c3 plants contributed significantly to the pom pool during the wet and dry period, respectively. based on isotopic mixing model, detrital om and phytoplankton mostly characterized the pom pools during the wet and dry periods, respectively. accordingly, our results suggest also that the zooplankton diet was mostly supported by detrital om during the wet period and by both phytoplankton and detrital om during the dry one. the zooplankton trophic level (tl, 2.7) during the wet period was relatively higher than that (1.9) during the dry one, suggesting a relative higher preference for detritus than phytoplankton during the wet period. the results of this study allowed us confirming that detrital om can significantly support zooplankton production in the godavari estuary. key words: phytoplankton, zooplankton, detritus, stable isotopes, river discharge, godavari estuary. received: december 2017. accepted: may 2018. no nco mm er cia l u se on ly j. mukherjee et al.2 web dynamics (post, 2002; finlay and kendall, 2007; hussey et al., 2014; middelburg, 2014). estuaries are among the most biogeochemically active zones on the surface of the earth (gattuso et al., 1998; cole et al., 2007; richey et al., 2002; bhavya et al., 2015) and higher rates of biological production are observed due to large nutrient inputs (nixon et al., 1986; kelly and levin, 1986; heip et al., 1995; caffrey, 2004; gazeau et al., 2004, 2005). estuaries receive a considerable amount of inorganic nutrients, organic matter, contaminants from the land and exchange them with the coastal ocean (nixon et al., 1986; howarth et al., 1996). besides the allochthonous sources, the active biological pump in estuaries produce significant amounts of organic matter that can accumulate both in the water column and sediments (wollast and mackenzie, 1989; smith and hollibaugh, 1993; wollast, 1998). while it is clear that these two organic matter sinks play a key role in the trophic dynamics of estuaries, their relative importance is unclear and may be regional specific due to variable composition and (nutritional) quality of organic matter (manini et al., 2003). within this framework, a significant amount of work was carried out in lakes. for instance, lammers et al. (2017) noticed that allochthonous organic matter can contribute significantly to bacterial diet during winter, whereas in summer and fall such a contribution is very low. strong evidence for the support of allochthonous organic matter to herbivorous zooplankton production was reported in several lakes (grey et al., 2001; brett et al., 2009; cole et al., 2011). suzuki et al. (2013), using stable isotopic composition of carbon, reported that the copepods diet in the chikugo river estuary, japan is supported by organic matter derived from phytoplankton ad their detrital remains. matson and brinson (1990), using stable isotopes of carbon, found that a significant fraction of zooplankton diet in the pamlico and neuse estuaries is contributed by terrestrial organic matter, and similar conclusions were made for some estuaries in siberia (doi et al., 2006). indian estuaries are characterized by runoff episodes associated with monsoonal precipitation (vijith et al., 2009; sridevi et al., 2015). the estuary behaves like a freshwater lake during the peak discharge period whereas seawater contribution increases during the dry period (sridevi et al., 2015). godavari is the largest monsoonal river in india and fed mainly by the southwest monsoonal precipitation during summer (june-september, wet period) (acharyya et al., 2012). during the wet period, freshwater brings large amounts of inorganic nutrients, organic matter and suspended matter to the estuary (sarma et al., 2009; 2010). during this period, despite the high nutrient concentrations, phytoplankton biomass is very low due to the high suspended load which limits light availability (acharyya et al., 2012). based on the values of the photosynthesis to respiration ratio, sarma et al. (2009) estimated that 40-60% and 70-95% of the heterotrophic production is supported by organic matter of terrestrial orifin during the wet and dry periods, respectively (gawade et al., 2017). the contribution of different taxa to the phytoplankton assemblages in the godavari estuary varies according to the magnitude of freshwater discharge rather than to nutrient concentrations (bharati et al., 2018). for instance, freshwater algae (e.g., chlorophyceae) dominate during the peak freshwater discharge period (july-august). a phytoplankton bloom (mainly due to cyanophyceae) occurs in the estuary once the river discharge decreases below 2000 m3 s–1 and the suspended matter load is <200 mg l–1 (bharati et al., 2018). when the freshwater discharge completely stops (january-may) due to drying of the upstream river and closing of dam gates, estuarine phytoplankton are dominated by bacillariophyceae (diatoms) (sarma et al., 2009, bharati et al., 2018). we hypothesized that the contrasting conditions between wet and dry periods and the associated variations in the composition of the potential food items may influence the zooplankton diet in the godavari estuary. to test this hypothesis, we investigated the sources of zooplankton diet in the godavari estuary during wet and dry periods using stable isotopic composition of carbon and nitrogen in the live and dead organic matter. methods study area godavari river is located between 16 and 18°n latitude and originates at an altitude of about 1600 m near nasik city in the western ghats. it flows eastwards across peninsular india for about 1480 km and drains into the bay of bengal at bhairavapalem, on the central east coast of india (fig. 1). godavari is the largest monsoonal river in india and has created an extensive delta on the east coast of india. the basin climate is generally dry with an average rainfall of 1512 mm y–1. the catchment receives about 82% of the annual rainfall during the summer and the rest in the winter monsoon (central pollution control board, 1995). the discharge of freshwater into the godavari estuary is controlled by century old low dam at dowleiswaram (fig. 1). discharge occurs between june and december with a peak in july-august and reduces considerably from october. after the dam, the river bifurcates into two major distributaries; the eastward flowing major tributary is called gautami-godavari, while the other flowing southwards is vasistha-godavari and the former is the major branch of river in terms of discharge. the present study was conducted in the gautami-godavari estuary at yanam during 2012-2013 (fig. 1). no nco mm er cia l u se on ly river discharge influence on zooplankton diet 3 sampling the samples were collected during november 2012 and january 2013, assumed to represent wet and dry conditions, respectively. at each sampling date, three water samples were collected at yanam, middle of the estuary using 5l niskin bottles operated onboard a hydrographic vessel, for the subsequent analyses of inorganic nutrients, phytoplankton biomass (in terms of chlorophyll-a chl-a concentrations), content and isotopic composition of carbon and nitrogen in pom. about 1 l of water sample filtered through pre-combusted gf/f filter for either chl-a concentrations or pom composition. an additional 1l of water sample was collected using plastic bottles for the taxonomic analysis of phytoplankton. zooplankton samples were collected by horizontal towing using a bongo net (200 µm pore size). chlorophyll-a analysis the chl-a retained on the filter was extracted with dimethyl formamide (dmf) at 4°c for 12 hours in the dark and the extract fluorescence was measured with a spectrofluorophotometer (varian instruments, palo alto, ca, usa) (suzuki and ishimaru, 1990). pom retained on the filter was first dried at 60°c overnight, and then kept in hcl acid fumes for 12 hours to remove inorganic carbon for the subsequent measurement of content and isotopic composition of carbon (δ13cpom). acid treatment was not done on filters dedicated to the analysis of the content and isotopic composition of nitrogen (δ15npom) (goering et al., 1990; bunn et al., 1995; pinnegar and polunin, 1999). phytoplankton and zooplankton taxonomy phytoplankton and zooplankton were separated under an upright microscope (4x magnification; olympus dx 53) with the help of an injection syringe and fine needle. the specimens were cleaned with milliq water and subsequently transferred to tin cups and dried at 60°c for 12 h. the dried tin cups were introduced to the elemental analyzer attached to the isotope ratio mass spectrometer (irms). inorganic nutrient analyses nutrients were analyzed following standard procedures (grashoff et al., 1992) following colorimetric method using auto analyzer (skalar san++, the netherlands). the analytical precision, expressed as standard deviation, was ±0.02, 0.02, 0.01 and 0.02 µm respectively for nitrate+nitrite, ammonium, phosphate and silicate. stable isotope analyses the content and isotopic ratios of carbon and nitrogen in the samples were measured using elemental analyzer (thermo electron, germany) coupled with isotope ratio mass spectrometer (irms delta v plus, finnigan, germany) through conflo iv interface, with oxidized column kept at 1050°c and the reduced one at 650°c. the results are expressed as relative to conventional standards, i.e., pdb for carbon (coplen, 1996) and atmospheric n2 for nitrogen (mariotti, 1983) as δ values, defined as: δr=[x sample xstandard)/ xstandard]*1000 (‰) (eq. 1) where, r=13c or 15n, and x=13c/12c or 15n/14n. highpurity co2 and n2 gases were used as working standards for carbon and nitrogen, respectively. these gases were calibrated with internal reference materials of glutamic acid, alanine and marine sediment and international standards obtained from the international atomic energy agency (iaea). the standard deviation on 20 aliquots of the same samples was < 0.2% for both δ13c and δ15n. the trophic level (tl) was estimated using δ15n isotopic values using the expression proposed by post (2002): tl= λ + (δ15nconsumer [δ15nbase1 x α + δ15nbase2 x (1α)])/3.4 (eq. 2) where, δ15nconsumer is the nitrogen isotopic ratio of the zooplankton. δ15nbase1 and δ15nbase2 are the isotopic composition of base 1 (phytoplankton) and base 1 (detritus), respectively. l is trophic position of the organism used to estimate δ15nbase and it is 1 for primary producers. α is the proportion of nitrogen derived from the base of food web one (base 1) to consumer (post,fig. 1. the study site. no nco mm er cia l u se on ly j. mukherjee et al.4 2002). when the movement of nitrogen and carbon through the food web is similar, α can be estimated using carbon isotopes α=(δ13cconsumer -δ13cbase2)/(δ13cbase1 δ13cbase2) (post, 2002). sources of the zooplankton diet the relative contribution of variable sources of food to consumers was quantified using isotope mixing model ‘siar’ (stable isotope analysis in r), an open source package (http://cran.r-project.org/), which resolves mixture within a bayesian framework (parnell et al., 2008, 2010). this model includes standard deviation as a residual error term to compute the variability (jackson et al., 2009; parnell et al., 2010). a detailed description of the model is given elsewhere (sarma et al., 2014). the selection of isotopic composition of the source is critical for the estimation of proportional contributions to consumers. mean isotopic ratios of δ13c and δ15n of several end members (c3 and c4 terrestrial plants, soil organic matter, fresh water algae, marine phytoplankton, estuarine phytoplankton, zooplankton) from drainage basin of the godavari river were measured (tab. 1) to separate and determine the percentage contribution of the live (autochthonous/in situ phytoplankton) and the detritral fractions (allochthonous/ex situ, dead organic matter) to the pom (krishna et al., 2015). the freshwater algae were separated under the microscope from water samples collected from the upstream godavari estuary (close to the dam) for fixing end members of in-situ sources. the end member values are close to those estimated elsewhere in other estuaries (hamilton and lewis, 1992; middelburg and nieuweuwenhuize, 1998; barth et al., 1998; hellings et al., 1999; kao and liu, 2000; boschker et al., 2005; bontes et al., 2006). results water column variables due to variable river discharge, hydrographic conditions varied considerably between the wet (junedecember) and dry (january-may) periods. during the wet period, the mean river discharge to the godavari estuary was relatively weak (~825 m3 s–1, on average), whereas no discharge at all occurred during the dry period (january). as a result of the weak discharge during the wet period, a strong stratification occurred in the estuary with low salinity waters (9.236) constrained at the surface and high salinity waters at the bottom (27.631). in contrast, during the dry period, high salinity (27.236) values characterized the entire water column. the highest concentrations of nutrients (nitrate, ammonium, phosphate and silicate of 10.9, 15.30, 2.91 and 52.82 µm respectively) were observed during the wet period, whereas values measured during the dry period were up to 5-10 times lower (1.8, 1.4, 0.9 and 10.3 µm, respectively). particulate organic c (poc) concentrations were significantly higher (t-test, t=245.1; p<0.001) during the wet period (3608±117 µgc l–1) than during the dry one (1284±67 µgc l–1). the c:n ratio values of pom were significantly (t-test; t=31.6; p<0.001) higher during the wet (19.4±3) than the dry period (14.6±2; tab. 1). phytoplankton abundance, biomass and assemblage composition phytoplankton abundance was lower (t-test; t=-70.4; p<0.001) during the wet period (44253 cell l–1) than during the dry one (63684 cell l–1; tab. 2), the phytoplankton biomass was significantly higher (t-test; t=25.8; p<0.001) during the wet period (10.5±2 µg chl-a l–1) than during the dry one (4.7±1 µg chl-a l–1) (tab. 2). values of the poc/chl-a ratio were higher during the wet (343) than the dry period (271) (t-test; t=34.6; p<0.001). during the wet period, the phytoplankton assemblage was dominated by cyanophyceae (mostly merismopedia sp. and gleocapsa sp.; overall 47% of the total phytoplankton abundance), followed by chlorophyceae (33%; mostly actinastrum sp., scenedesmus sp., pediastrum sp.), small (<10 µm) diatoms (19%, leptocylindrus sp., coscinodiscus sp.) and dinophyceae (1%). during the dry period small size diatoms (mostly coscinodiscus sp., chaetoceros sp. and ceratium sp.) represented up to 95% of the total phytoplankton abundance. tab. 1. the isotopic composition of different sources of organic matter used in the siar model. source δ13c (‰) δ15n (‰) reference c3 plants -25.9±1.2 5.1±2.1 krishna et al., 2015 c4 plants -13.1±1.2 4.4±2.1 krishna et al., 2015 marine phytoplankton -23.6±0.3 9.2±0.8 krishna et al., 2015 soil organic matter -19.2±2.4 10.3±2.7 krishna et al., 2015 freshwater algae -33.1±2.3 6.8±0.8 this study no nco mm er cia l u se on ly river discharge influence on zooplankton diet 5 zooplankton abundance and community composition mesozooplankton abundance was relatively lower during the dry (105 ind. m–3) than the wet period (188 ind. m–3). during the wet period the mesozooplankton assemblage consisted mostly of copepods (85% of the total zooplankton abundance) and other, less abundant, taxa were cladocera, gastropod veliger, nauplii, tellina sp, and zoea (cumulatively representing 15% of the total abundance). calanoida represented up to 90% of the total zooplankton abundance, followed by cyclopoida (5%), harpacticoida (3%) and polycheate larvae (2%). during the dry period, copepods (calanoida, cyclopoida and harpacticoida) were the most abundant (92% of the total mesozooplankton abundance), followed by zoea, decapoda, and harpacticoid copepoda (cumulatively 8%). isotopic composition of phytoplankton and zooplankton the isotopic composition of carbon in the different phytoplankton taxa ranged between -28.1 and -24.9‰ and a relatively higher value characterized cyanophyceae (-28.1‰) during the wet period (tab. 3). the mean δ13c (δ15n) of phytoplankton was significantly lower (higher) (t-test; t=-8.9; p<0.01 and t=53.21; p<0.01, respectively) during the wet (-26.1±1.4‰ and 15.9±2.5‰ respectively) than the dry period (24.6±0.6‰ and 4.6±1.0‰ respectively) (tab. 2). the isotopic composition of δ13c and δ15n of mesozooplankton ranged from -29.3 to -29.1‰ and 13.9 to 15.2‰, respectively during wet period (tab. 3), and from -23.8 to -22.2‰ and 6.9 to 8.4‰ during the dry one (tab. 3). the δ13c (δ15n) of mesozooplankton during the dry period was significantly higher (lower) during the wet one (t-test; t=-44.3; p<0.001 and t=140.5; p<0.001, respectively; tab. 2). sources and isotopic composition of pom and detritus the δ13c and δ15n of pom was relatively lighter during the wet (-31.2‰ and 4.9‰) than the dry period (25.6‰ and 6.4‰). the results obtained from the siar model suggest that 17 and 25% of pom was contributed by live organic matter during the wet and dry periods, respectively (tab. 4). during the wet period freshwater algae contributed the most 69%) to detritus, followed by c3 plants (9%), whereas during the dry periods c3 plants (60%) were the most important contributors to detritus, followed by estuarine phytoplankton (14%) and freshwater algae (12%) (tab. 4). during the wet period the isotopic composition of carbon in detritus (δ13cdet), resembling that of freshwater algae, was significantly (t-test; t=-56.5; p<0.001) depleted (-31.7±1.4‰), when compared to that measured during the dry one (-25.9±1.3‰; tab. 2). the isotopic composition of nitrogen in detritus (δ15ndet) was significantly higher (t-test; t=5.9; p<0.01) during the wet period (6.5±3.7‰) than the dry one (5.3±1.3‰; tab. 2). trophic levels and food web structure during the wet period, due to the higher δ15nphytoplankton, the δ15nbase values for consumers were enriched (10.3 and 7.5‰) when compared to those estimated during the dry one (4.9 and 6.2‰ respectively). during the wet period, the estimated tl for calanoida and cyclopoida was 2.6 and 2.7, respectively and decreased (1.8 and 2.0, respectively) during the dry period (tab. 5). the mean tl tab. 2. mean concentrations and isotopic values of several components of the godavari estuary ecosystem during wet and dry. property wet period dry period t-test and p-value discharge (m3 s–1) 825±80 0 142.04; <0.001 chlorophyll-a (µg l–1) 10.5±2 4.7±1 25.88; <0.001 pom (µgc l–1) 3608±117 1284±67 245.15; <0.001 pon (µgn l–1) 238±3.2 210±2.9 11.17; <0.001 c/n 19.4±3 14.6±2 31.64; <0.001 phytoplankton abundance (cells/l) 44253 63684 -70.40; <0.001 poc/chl 343±18 271±12 34.60; <0.001 dominant group cyanophyceae bacillariophyceae — δ13cdetritus (‰) -31.7±1.4 -25.9±1.3 -56.51; <0.001 δ15ndetritus (‰) 6.5±3.7 5.3±1.3 5.87; <0.01 δ13cphytoplankton (‰) -26.1±1.4 -24.6±0.6 -8.96; <0.01 δ15nphytoplankton (‰) 15.9±2.5 4.6±1.0 53.21; <0.001 δ13czooplankton (‰) -29.1±0.1 -23.3±0.8 -44.31; <0.001 δ15nzooplankton (‰) 14.5±0.6 7.9±0.7 140.48; <0.001 no nco mm er cia l u se on ly j. mukherjee et al.6 for zooplankton was 2.7 and 1.9 during the wet and dry periods, respectively. discussion variable sources of pom during the wet and dry periods the carbon to nitrogen (c:n) ratio is a trace to identify source of organic matter as it varies for variable sources such as plankton (6-7), bacteria (4-5), and organic matter from higher plants (>20) (hedges et al., 1997). however, several diagenetic processes can modify c:n ratios which lower its viability to identify the actual sources of organic matter. for instance, c:n ratio of higher plant litter typically decreases due to bacterial colonization, and increases in senescent or dead algae due to the preferential removal of nitrogen by consumers (hedges et al., 1997; herman and heip, 1999). nonetheless, though such biases do not allow identifying the exact source of organic matter, variations in the c:n ratio values can provide some indication on changes occurring in om origin and tab. 3. isotopic composition of sources, primary, secondary and tertiary consumers and trophic level of primary and secondary consumers during the wet and dry periods. source/consumer groups δ13c (‰) δ15n (‰) tl wet period source phytoplankton* -26.1±0.9 15.9±2.5 detritus -31.7±1.4 6.5±3.7 soil organic matter -23.5±1.1 6.0±0.9 zooplankton calanoida (10) -29.3±1.2 13.9±2.5 2.6 cyclopoida (21) -29.1±0.6 14.5±3.4 2.7 cladocera (11) -29.2±0.5 15.2±0.3 2.9 dry period source phytoplankton (100) -24.6±0.4 4.6±1.0 detritus -25.9±1.2 5.3±1.3 soil organic matter -23.1±0.8 6.3±0.5 zooplankton harpacticoida (14) -22.5±0.3 7.4±0.6 1.8 cyclopoida (25) -24.2±0.6 6.9±0.8 1.6 zoea (25) -22.2±0.5 8.0±0.5 2.0 calanoida (25) -23.8±0.5 8.2±0.3 2.0 decapoda (16) -23.4±0.8 8.1±0.4 2.0 tl, trophic level; *nearly 100 number of phytoplankton cells were analyzed. tab. 4. proportion (%) of source contributors to the particulate organic matter pool in the godavari estuary based on the outputs of the siar model. pom period wet period dry period live fraction phytoplankton 7 14 zooplankton 8 7 marine algae 3 3 detrital fraction freshwater algae 69 12 c3 plants 9 60 c4 plants 4 4 pom, particulate organic matter. tab. 5. contribution (%) of detritus and phytoplankton to the zooplankton diet in the godavari estuary based on siar model. consumers period detritus phytoplankton zooplankton wet 60 40 zooplankton dry 40 60 no nco mm er cia l u se on ly river discharge influence on zooplankton diet 7 diagenetic status (middelburg and herman, 2007). the higher values of the c:n ratio values of pom observed during the wet period (tab. 1) suggest that during this period the godavari estuary was characterized by a relevant fraction of non-living organic matter. the poc:chl-a ratio (c:chl-a) can also be used to delineate the sources of organic matter as relatively lower values are associated with freshly derived organic matter by in situ biological production, whereas higher values are associated with older and degraded om (cifuentes et al., 1988; richard et al., 1997; bentaleb et al., 1998). in fact, the c:chl-a ratio of fresh organic matter produced by marine phytoplankton varies from ~40 (montagnes et al., 1994), <70 (geider, 1987) <100 (head et al., 1996), <140 (thompson et al., 1992) to <200 (cifuentes et al., 1988; bentaleb et al., 1998) also according to regional temperature and irradiance regimes, as well as species growth rates and composition (heath et al., 1990; montagnes et al., 1994; geider et al., 1998). in our study, the c:chl-a ratio during the wet (343±18) and dry (271±12) period were consistently >200, suggesting the presence in both periods of a relevant fraction of terrestrial om sources, more evidently during the wet period (tab. 2), as previously reported from other indian estuaries (sarma et al., 2014). the output of the siar model suggests that 18 and 24% (wet and dry period, respectively) of the pom was contributed by live om, also that the contribution of detrital om was high during both study periods (tab. 4). such low contribution of live om was caused by the minimal photic depth (0.1 to 2 m; sarma et al., 2009) which likely limited primary production. during the wet period, the δ13cdetritus (-31.7±1.4 ‰) is close to that of freshwater algae (-33.2‰; tab. 1), suggesting that dead freshwater algae contributed significantly to the detritus pool. during the wet period, apart from estuarine phytoplankton, the estuary received organic matter also from terrestrial sources, such as c3, c4 land plants and soil om (sarma et al., 2014). moreover, the output of the siar model suggests that, during the wet period, 69% of the detritus was contributed by freshwater algae and 9% by c3 plants (tab. 4). the accumulation of detritus in the water column depends also upon the residence time of water in the estuary. the residence time of water in the godavari estuary is <1 d during the peak discharge and increases up to >30 d during the dry period. however, also when a moderate discharge occurs (november, wet period) the residence time of estuarine water is >20 d (sridevi et al., 2015). moreover, during november, freshwater phytoplankton blooms have been also reported in the dowleiswaram dam reservoir waters (prasad et al., 2013), which could have been injected into the estuary along with freshwater discharge during the wet period. the combination of varying residence times and the potential input of waters from the dam can be reasonably invoked to explain the relative importance of dead freshwater contribution to the detritus pool during the wet period. on the other hand, during the dry period, ~60% of detritus was contributed by c3 plants and 12% by freshwater algae. despite the negligible amount of discharge during the dry period a certain amount of terrestrial om, brought during discharge period, might have been trapped in the estuary due to high residence time of water and recirculation through tidal mixing. thus, the results of our study suggest that during both wet and dry periods the contribution of in situ phytoplankton to pom in the godavari estuary is very small when compared to allochthonous om sources, like terrestrial om or freshwater algae. potential sources of nutrients for phytoplankton the δ15nphytoplankton was significantly heavier (15.9±2.5 ‰) during wet period than earlier reports from the godavari estuary (5.1 to 7.8 ‰; sarma et al., 2012, 2014). such heavier δ15nphytoplankton were normally observed in highly polluted estuaries (up to 23‰; middelburg and herman, 2007; kromkamp et al., 1995) and were at times attributed to the high chemoautotrophic production rates by nitrifiers (soetaert and herman, 1995a, 1995b). owens (1985) noticed enriched δ15n of pn (14.7‰) in the suspended matter at the turbidity maximum zone and attributed it to intense biological processing of om. seasonal enrichment of δ15npom (18-24‰) has been also reported repeatedly during spring in several other estuaries (middelburg and herman, 2007), and attributed to the utilization of isotopically enriched nitrogen, especially residual ammonium resulted from nitrification (mariotti et al., 1984) or extensive algal uptake of nitrogen, leading to enrichment of leftover nitrogen. the occurrence of phytoplankton blooms associated with rapid decrease in din concentrations was reported a month prior to our sampling (sarma et al., 2009). based on previous studies conducted in other estuaries, we can hypothesize that in our study δ15ndin might be enriched in october due to the extensive utilization of lighter inorganic nitrogen available in association with the phytoplankton bloom and that the uptake of such enriched δ15ndin might have increased isotopic value of phytoplankton during the wet period. though the detritus pool includes phytoplankton biomass, during the dry period the δ15npom were lower than δ15nphytoplankton suggesting that the contribution of the latter may be less than the former. on the other hand, during the dry period, δ15nphytoplankton (4.6±1.0‰) was close to that of the nutrients derived from regeneration of marine organic matter (4.8‰; sigman et al., 2000) suggesting that regenerated nutrients might have supported phytoplankton biomass during the dry period. no nco mm er cia l u se on ly j. mukherjee et al.8 potential sources for the zooplankton diet the output of the siar model suggests that zooplankton based for their diet preferentially (60%) on detritus than phytoplankton (40%) during the wet period, and inverted such preference during the dry one (tab. 5). despite during the wet period higher phytoplankton biomass was observed (10.5±2 µg chl-a l–1) zooplankton preferred detritus as a food source. in this regard, however, it is worth noting that the detritus pool during the wet period contained an important fraction of freshwater dead/senescent algae, likely providing a labile source of food as the one provided by phytoplankton biomass. during the dry period, instead, the preference of zooplankton for phytoplankton (60%) depended most likely by the availability of larger size phytoplankton. trophic level of food web during wet and dry periods the computed tl for zooplankton (calanoida, and cyclopoida) was 2.6-2.9 during the wet period and decreased (1.6-2.0) during the dry one (tab. 3). such a difference can be attributable to variations in the base of the food web. in fact, during the dry period, when phytoplankton represented a relevant proportion of the zooplankton diet, the tl for primary consumers was low, whereas the more important contribution of detritus during the wet period resulted in a higher tl (tab. 3). these results are also consistent with the variations in the relative importance of detritus vs phytoplankton to the zooplankton diet between the two sampling periods. variations in the relative importance of detritus (40% vs 60% in the dry and wet periods, respectively), were also associated with changes in the relative importance of detritus from c3 plants (prevailing during the dry period) and freshwater algae (prevailing in the wet period) (tab. 4). conclusions our results confirm previous findings showing that mesozooplankton can modify their tl in response to natural environmental changes, resulting, in turn, in expansions or contractions in trophic linkages within the food web and, as a consequence, affecting the efficiency of energy transfer in food webs (decima et al., 2013). for instance, landry (1981) reported that several species may alter their dietary compositions and tl within the food web as a consequence of changes in the size structure and availability of phytoplankton with a preference for larger cells (frost, 1972; landry, 1981; ohman and runge, 1994), either as a passive response to relative availability of alternate prey or an active switching tendency toward omnivory, when mean phytoplankton size is smaller (calbet and landry, 1999). as changes in tl of primary consumers affect, by cascade, the trophic position of consumers at higher fig. 2. the isotopic composition of carbon and nitrogen of phytoplankton, detritus and zooplankton during wet (a) and dry (b) periods in the godavari estuary. no nco mm er cia l u se on ly river discharge influence on zooplankton diet 9 trophic levels, our results, based on the estimates of the trophic position through δ15n isotopic values, suggest the presence of a food chain that is longer during the dry than the wet period (fig. 2). although based on an under-replicated study, we conclude that changes in the structure (length) of the food chain in the godavari estuary reported here were plausibly determined by variations in the array of environmental factors, which, in turn, affected the origin, composition and food availability of pom at the base of the food web and, by cascade, the zooplankton diet. acknowledgments we thank the director, national institute of oceanography (nio), goa, the scientist-in-charge, nioregional centre, visakhapatnam and dr. a.c. anil, du leader, for their support and encouragement. the work was carried out under the council of scientific and industrial research (csir), government of india funded project (ocean finder, psc 0105). we also thank the boat personnel for their help during field trip. we would like to acknowledge mr. p. praveen kumar and g. srikanth for their help during measurements on irms. we would like to thank two anonymous reviewers and editor-in-chief for their constructive criticism to improve presentation of this manuscript. references acharyya t, sarma vvss, sridevi b, venkataramana v, bharathi md, naidu sa, kumar bsk, prasad vr, bandopadhyay d, reddy npc, kumar md, 2012. reduced river discharge intensifies phytoplankton bloom in godavari estuary, india. mar. chem. 132-133:15-22. barth jac, veizer j, mayer b, 1998. origin of particulate organic carbon in the upper st.lawrence: isotopic constraints. earth planet. sci. lett. 162:11-121. bentaleb i, fontugne m, descolas-gros c, girardin c, mariotti a, pierre c, brunet c, poisson a, 1998. carbon isotopicfractionation by plankton in the southern indian ocean: relation between δ13c of particulate organic carbon and dissolved carbon dioxide. j. marine syst. 17:39-58. bharati md, sarma vvss, ramaneswari k, 2018. intra-annual variations in phytoplankton biomass and its composition in the tropical estuary: influence of river discharge. mar. pollut. bull. 129:14-25. bhavya ps, kumar s, gupta gvm, sudheesh v, sudharma kv, varrier, ds, dhanya, kr, saravanane n, 2015. nitrogen uptake dynamics in a tropical eutrophic estuary (cochin, india) and adjacent coastal waters. estuar. coast. 39:4-67. bontes bm, pel r, ibelings bw, boschker hts, middelburg jj, van donk e, 2006. the effects of biomanipulation on the biogeochemistry, carbon isotopic composition and pelagic food web relations of a shallow lake. biogeosciences 3:69-83. boschker hts, middelburg jj, kromkamp j, 2005. biomarker and carbon isotopic constraints on bacterial and algal community structure and functioning in a turbid, tidal estuary. limnol. oceanogr. 50:70-80. brett mt, kainz mj, taipale sj, seshan h, 2009. phytoplankton, not allochthonous carbon, sustains herbivorous zooplankton production. p. natl. acad. sci. usa 106:21197-21201. bunn se, loneragan nr, kempster ma, 1995. effects of acid washing on stable isotope ratios of c and n in penaeid shrimp and seagrass: implications for food-web studies using multiple stable isotopes. limnol. oceanogr. 40:622-625. caffrey jm, 2004. factors controlling net ecosystem metabolism in us estuaries. estuaries 27:90-101. calbet a, landry mr, 1999. mesozooplankton influences on the microbial food web: direct and indirect trophic interactions in the oligotrophic open ocean. limnol. oceanogr. 44:1370-1380. central pollution control board, 1995. basin sub-basin inventory of water pollution godavari basin. cpcb, delhi. cifuentes aa, sharp jh, fogel ml, 1988. stable carbon and nitrogen isotope biogeochemistry in the delaware estuary. limnol. oceanogr. 33:1102-1115. cole jj, prairie yt, caraco nf, mcdowell wh, tranvik lj, striegl, rg, duarte cm, kortelainen p, downing ja, middelburg jj, melack j, 2007. plumbing the global carbon cycle: integrating inland waters into the terrestrial carbon budget. ecosystems 10:171-184. cole jj, carpenter sr, kitchell jf, pace ml, solomon ct, weidel bc, 2011. strong evidence for terrestrial support of zooplankton in small lakes based on stable isotopes of carbon, nitrogen, and hydrogen. p. natl. acad. sci. usa 108:1975-1980. coplen tb, 1996. new guidelines for reporting stable hydrogen, carbon and oxygen isotope-ratio data. geochim. cosmochim. acta 60:3359-3360. cushing dh, 1975. marine ecology and fisheries. cambridge university press, cambridge: 278 pp. decima m, landry mr, popp b, 2013. environmental perturbation effects on baseline d15n values and zooplankton trophic flexibility in the southern california current ecosystem. limnol. oceanogr. 58:624-634. deniro mj, epstein s, 1978. influence of diet on the distribution of carbon isotopes in animals. geochim. cosmochim. acta 42:495-506. deniro mj, epstein s, 1981. influence of diet on the distribution of nitrogen isotopes in animals. geochim. cosmochim. acta 45:341-351. doi h, zuykova ei, kikuchi e, shikano s, kanou k, yurlova n, yadrenkina e, 2006. spatial changes in carbon and nitrogen stable isotopes of the plankton food web in a saline lake ecosystem. hydrobiologia. 571:395-400. finlay jc, kendall c, 2008. stable isotope tracing of temporal and spatial variability in organic matter sources to freshwater ecosystems, p. 283-333. in: r. michener and k. lajtha (eds.), stable isotopes in ecology and environmental sciences. blackwell publishing ltd., malden. france r, 1995. critical examination of stable isotope analysis as a means for tracing carbon pathways in stream ecosystems. can. j. fish. aquat. sci. 52:651-656. frost bw, 1972. effects of size and concentration of food no nco mm er cia l u se on ly j. mukherjee et al.10 particles on feeding behavior of marine planktonic copepod calanus pacificus. limnol. oceanogr. 17:805-815. gattuso jp, frankignoulle m, wollast r, 1998. carbon and carbonate metabolism in coastal aquatic ecosystems. ann. rev. ecol. evol. s. 29: 405-434. gawade l, sarma vvss, rao yv, hemalatha kpj, 2017. variation of bacterial metabolic arates and organic matter in the monsoon-affected tropical estuary (godavari, india). geomicrobiol. j. 34:628-640. gazeau f, smith sv, gentili, b, frankignoulle m, gattuso jp, 2004. the european coastal zone: characterization and first assessment of ecosystem metabolism. estuar. coast. shelf sci. 60:673-694. gazeau f, borges av, barron c, duarte cm, iversen n, middelburg jj, delille b, pizay md, frankignoulle m, gattuso jp, 2005. net ecosystem metabolism in a micro-tidal estuary (randers fjord, denmark): evaluation of methods. mar. ecol. prog. ser. 301:23-41. geider rj, 1987. light and temperature dependence of the carbon to chlorophyll-a ratio in microalgae and cyanobacteria: implications for physiology and growth of phytoplankton, new phytol. 106:1-34. geider rj, mcintyre hl, kana tm, 1998. a dynamic regulatory model of phytoplanktonic acclimation to light, nutrients, and temperature, limnol. oceanogr. 43:679-694. goering j, alexander v, haubenstock n, 1990. seasonal variability of stable carbon and nitrogen isotope ratios of organisms in a north pacific bay. estuar. coast. shelf s. 30:239-260. grashoff k, ehrhardt m, kremling k, 1992. methods of seawater analysis. verlag chemie, new york: 419 pp. grey j, jones ri, 2001. seasonal changes in the importance of the source of organic matter to the diet of zooplankton in loch ness, as indicated by stable isotope analysis. limnol. oceanogr. 46:505-513. hall sj, raffaelli dg, 1993. food webs: theory and reality. adv. ecol. res. 24:187-239. hamilton sk, lewis wm, 1992. stable carbon and nitrogen isotopes in algae and detritus from the orinoco river flood plain, venezuela. geochim. cosmochim. acta 56:42374246. head ejh, harrison wg, irwin bi, horne epw, li wkw, 1996. plankton dynamics and carbon flux in an area of upwelling off the coast of morocco. deep-sea res. pt i 43:1713-1738. heath mr, richardson k, kiørboe t, 1990. optical assessment of phytoplankton nutrient depletion. j. plankton res. 12:381-396. hedges ji, keil rg, benner r, 1997. what happens to terrestrial organic matter in the ocean? org. geochem. 27:195-212. herman pmj, heip chr, 1999. biogeochemistry of the maximum turbidity zone of estuaries (mature): some conclusions. j. marine syst. 22:89-104. heip chr, goosen nk, herman pmj, kromkamp j, middelburg jj, soetaert k, 1995. production and consumption of biological particles in temperate tidal estuaries. oceanogr. mar. biol. 33: 1-149. hellings l, dehairs f, tackx m, keppens e, baeyens w, 1999. origin and fate or organic carbon in the freshwater part of the scheldt estuary as traced by stable carbon isotope composition. biogeochemistry 47:167-186. hinz h, moranta j, balestrini s, sciberras m, pantin, jr, monnington j, zalewski a, kaiser mj, skold m, jonsson p, bastardie f, hiddink jg. 2017. stable isotopes reveal the effect of trawl fisheries on the diet of commercially exploited species. sci. rep. 7:6334. hobson ka, clark rg, 1992. assessing avian diets using stable isotopes i: turnover of 13c in tissues. condor 94:181-88. hobson ka, fisk a, karnovsky n, holst m, gagnon jm, fortier m, 2002. a stable isotope (δ13c, δ15n) model for the north water food web: implications for evaluating trophodynamics and the flow of energy and contaminants. deep-sea res. pt. ii 49:5131-5150. howarth rw, schneider wr, swaney d, 1996. metabolism and organic carbon fluxes in the tidal freshwater hudson river. estuaries 19:848-865. hussey ne, macneil ma, mcmeans bc, olin ja, dudley sfj, cliff g, wintner sp, fenessy, st, fisk at, 2014. rescaling the trophic structure of marine food webs. ecol. lett. 17:239-250. jennings s, pinnegar jk, polunin nvc, warr kj, 2002. linking size-based and trophic analyses of benthic community structure. mar. ecol. prog. ser. 226:77-85. jackson al, inger r, bearhop s, parnell a, 2009. erroneous behaviour of mixsir, a recently published bayesian isotope mixing model: a discussion of moore & semmens. ecol. lett. 12: e1-e5. kao sj, liu kk, 2000. stable carbon and nitrogen isotope systematic in a human disturbed watershed (lanyang-his) in taiwan and the estimation of biogenic particulate organic carbon and nitrogen fluxes. global biogeochem. cy. 14:189-198. kelly jr, levin sa, 1986. a comparison of aquatic and terrestrial nutrient cycling and production processes in natural ecosystems, with reference to ecological consequences to some waste disposal issues, p. 165-203. in: g. kullenberg (ed.), the role of oceans as a waste disposal option. nato advanced research workshop series. d. reidel publ., dordrecht. krishna ms, naidu sa, subbaiah chv, gawade l, sarma vvss, reddy npc, 2015. sources, distribution and preservation of organic matter in a tropical estuary (godavari, india). estuaries coasts 38:1032-1047. kromkamp j, peene j, rijswijk p, van sandee a, goosen n, 1995. nutrients, light and primary production by phytoplankton and microphytobenthos in the eutrophic, turbid westerschelde estuary (the netherlands). hydrobiologia 311: 9-19. landry mr, 1981. switching between herbivory and carnivory by the planktonic marine copepod calanus pacificus. mar. biol. 65:77-82. lammers jm, reichart gj, middelburg jj, 2017. seasonal variability in phytoplankton stable carbon isotope ratios and bacterial carbon sources in a shallow dutch lake. limnol. oceanogr. 62:2773-2787. layman ca, arauj́o ms, boucek r, hammerschlag-peyer cm, harrison e, jud zr, matich p, rosenblatt ae, vaudo jj, yeager la, post dm, bearhop s, 2012. applying stable isotopes to examine food-web structure: an overview of analytical tools. biol. rev. 87:545-562. no nco mm er cia l u se on ly river discharge influence on zooplankton diet 11 manini e, fiordelmondo c, gambi c, pusceddu a, danovaro r, 2003. benthic microbial loop functioning in coastal lagoons: a comparative approach. oceanol. acta 26:27-38. mariotti a, 1983. atmospheric nitrogen is a reliable standard for natural 15n abundance measurements. nature 303:685-687. mariotti a, lancelot c, billen g, 1984. natural isotopic composition of nitrogen as a tracer of origin for suspended matter in the scheldt estuary. geochim. cosmochim. acta 48:549-555. matson ea, brinson mm, 1990. stabile carbon isotopes and the c:n ratio in the estuaries of’ the pamlico and neuse rivers, north carolina. limnol. oceanogr. 35:1290-1300. matthews b, mazumder a, 2003. compositional and inter-lake variability of zooplankton affect baseline stable isotope signatures. limnol. oceanogr. 48:1977-1987. matthews b, mazumder a, 2006. habitat specialization and the exploitation of allochthonous carbon by zooplankton. ecology 87:2800-2812. meili m, kling gw, fry b, bell rt, 1996. sources and partitioning of organic matter in a pelagic microbial food web inferred from the isotopic composition (δ13c and δ15n) of zooplankton species. arch. hydrobiol. special issues adv. limnol. 48:53-61. middelburg jj, 2014. stable isotopes dissect aquatic food webs from the top to the bottom. biogeosciences 11:2357-2371. middelburg jj, herman pmj, 2007. organic matter processing in tidal estuaries. mar. chem. 106:127-147. middelburg jj, nieuweuwenhuize j, 1998. carbon and nitrogen stable isotopes in suspended matter and sediments from the schelde estuary. mar. chem. 60: 217-225. minagawa m, wada e, 1984. stepwise enrichment of 15n along food chains: further evidence and the relation between 15n and animal age. geochim. cosmochim. acta 48:1135-1140. montagnes djs, berges ja, harrison pj, taylor fjr, 1994. estimating carbon, nitrogen, protein and chlorophyll a from volume in marine phytoplankton. limnol. oceanogr. 39:1044-1060. nixon sw, oviatt ca, frithsen j, sullivan b, 1986. nutrients and the productivity of estuarine and coastal marine ecosystems. j. limnol. soc. south africa 12:43-71. ohman md, runge ja, 1994. sustained fecundity when phytoplankton resources are in short supply: omnivory by calanus finmarchicus in the gulf of st lawrence. limnol. oceanogr. 39:21-36. owens njp, 1985. variations in the natural abundance of 15n in estuarine suspended particulate matter: a specific indicator of biological processing. estuar. coast. shelf sci. 20: 505510. parnell a, jackson a, 2008. siar: stable isotope analysis in r. r package version 3.3, http://cran.r-project.org/web/ packages/siar/siar.pdf parnell ac, inger r, bearhop s, jackson al, 2010. source partitioning using stable isotopes: coping with too much variation. plos one 5:e9672. peterson bj, fry b, 1987. stable isotopes in ecosystem studies. annu. rev. ecol. syst. 18:293-320. pinnegar jk, polunin nvc, 1999. differential fractionation of δ13c and δ15n among fish tissues: implications for the study of trophic interactions. functional ecol. 13:225-231. post dm, 2002. using stable isotopes to estimate trophic position: models, methods, and assumptions. ecology 83:703-718. prasad mhk, sarma vvss, sarma vv, krishna ms, reddy npc, 2013. carbon dioxide emissions from the tropical dowleiswaram reservoir on the godavari river, south east of india. j. water res. prot. 5:534-545. pope jg, shepherd jg, webb j, 1994. successful surf-riding on size-spectra: the secret of survival in the sea. philos. t. r. soc. b 343:41-49. richard, p, riera p, galois r, 1997. temporal variations in the chemical and carbon isotope compositions of marine and terrestrial organic inputs in the bay of marennes-oléron, france. j. coastal res. 13:879-899. richey je, melack jm, aufdenkampe ak, ballester vm, hess ll, 2002. outgassing from amazonian rivers and wetlands as a large tropical source of atmospheric co2. nature 416:617-620. santer b, sommerwerk n, grey j, 2006. food niches of cyclopoid copepods in eutrophic plubsee determined by stable isotope analysis. arch. hydrobiol. 167:301-316. sarma vvss, arya j, subbaiah chv, naidu sa, gawade l, praveen kumar p, reddy npc, 2012. stable isotopes of carbon and nitrogen in suspended matter and sediments from the godavari estuary. j. oceanogr. 68:307-319. sarma vvss, gupta snm, babu pvr, acharya t, harikrishnachari n, vishnuvardhan k, rao ns, reddy npc, sarma vv, sadhuram y, murty tvr, kumar md, 2009. influence of river discharge on plankton metabolic rates in the tropical monsoon driven godavari estuary, india. estuar. coast. shelf s. 85:515-524. sarma vvss, krishna ms, prasad vr, kumar bsk, naidu sa, rao gd, viswanadham r, sridevi t, kumar pp, reddy npc, 2014. distribution and sources of particulate organic matter in the indian monsoonal estuaries during monsoon. j. geophys. res. 119:2095-2111. sarma vvss, prasad vr, kumar bsk, rajeev k, devi bmm, reddy npc, sarma vv, kumar md, 2010. intra-annual variability in nutrients in the godavari estuary, india. cont. shelf res. 30:2005-2014. sigman dm, altabet ma, mc corkle dc, francois r, fischer g, 2000. the δ15n of nitrate in the southern ocean: nitrogen cycling and circulation in the ocean interior. j. geophys. res. 105:19599-19614. smith sv, hollibaugh jt, 1993. coastal metabolism and the oceanic organic carbon balance. rev. geophys. 31:75-89. soetaert k, herman pmj, 1995a. carbon flows in the westerschelde estuary (the netherlands) evaluated by means of an ecosystem model (moses). hydrobiologia 311:247-266. soetaert k, herman pmj, 1995b. nitrogen dynamics in the westerschelde estuary (sw netherlands) estimated by means of the ecosystem model moses. hydrobiologia 311:225-246. sridevi b, sarma vvss, murty tvr, sadhuram y, reddy npc, vijayakumar k, raju nsn, jawahar kumar ch, raju ysn, luis r, kumar md, and prasad kvsr, 2015. variability in stratification and flushing times of the gautami-godavari estuary, india. j. earth sys. sci. 124:993-1003. stowasser g, atkinson a, mcgill rar, phillips ar, collins ma, pond dw, 2012. food web dynamics in the scotia sea no nco mm er cia l u se on ly j. mukherjee et al.12 in summer: a stable isotope study. deep-sea res. pt. ii 5960:208-221. suzuki kw, ueda h, nakayama k, tanaka m, 2013. spatiotemporal dynamics of stable carbon isotope ratios in two sympatric oligohaline copepods in relation to the estuarine turbidity maximum (chikugo river, japan): implications for food sources. j. plankton res. 36:461-474. suzuki r, ishimaru t, 1990. an improved method for the determination of phytoplankton chlorophyll using n, ndimethylformamide. j. oceanogr. 46:190-194. thompson pa, guo mx, harrison pj, 1992. effects of variation in temperature. i. on the biochemical composition of eight species of marine phytoplankton. j. phycol. 28:481-488. van oevelen d, soetaert k, heip chr, 2012. carbon flows in the benthic food web of the porcupine abyssal plain: the (un)importance of labile detritus in supporting microbial and faunal carbon demands. limnol. oceanogr. 7:645-664. vander-zanden mj, cabana g, rasmussem jb, 1997. comparing trophic position of freshwater fish calculated using stable nitrogen isotope ratios (δ15n) and literature dietary data. can. j. fish aquat. sci. 54:1142-1158. vijith vd, sundar d, shetye sr, 2009. time-dependence of salinity in monsoonal estuaries. estuar. coast. shelf s. 85:601-608. wollast r, 1998. evaluation and comparison of the global carbon cycle in the coastal zone and in the open ocean, p. 213-252. in: k.h. brink and a.r. robinson (eds.), the sea: the global coastal ocean. processes and methods. j. wiley & sons, new york. wollast r, mackenzie ft, 1989. global biogeochemical cycles and climate, p. 453-473. in: a. berger (ed.), climate and geo-sciences. kluwer academic publishers, dordrecht. no nco mm er cia l u se on ly layout 1 introduction multiple and intersecting evolutionary processes are at the base of assembly and functioning of ecological communities (hendry, 2016). the latter are complex entities including populations of species whose ecological roles stem from several overlapping natural history processes, in which casualties and environmental constraints play simultaneously (gould, 2002; koonin, 2011). understanding the evolutionary history of species, the possible drivers of species life-histories, the ecological benefit of inter-specific interactions and, ultimately, evolutionary processes behind biodiversity are of pivotal importance for ecosystem studies (levin, 2007; hendry, 2016). conceptual and methodological approaches intersecting ecology and evolution are frequently applied to study plankton, a community of rapidly evolving and strongly interconnected species including both unicellular and multicellular organisms (lima-mendez et al., 2015; d’alelio et al., 2016a). the huge genetic diversity of plankton provides a molecular basis to an overwhelming phenotypic variability (de vargas et al., 2015; sunagawa et al., 2015). for instance: plankton individual-sizes span three orders of magnitude (boyce et al., 2015); morphological characteristics, like surface-to-volume ratio, are extremely variable even within a single aquatic system (morabito et al., 2007); coloniality is wide-spread among distantly related phyla (e.g., from diatoms to pelagic tunicates; bone and others, 1998; seckbach and kociolek, 2011); mixotrophy, or the contemporary presence of heterotrophic and autotrophic metabolism within the same organism, is common in planktonic protists (stoecker et al., 2017); several intersecting trophic interactions may establish among plankters (d’alelio et al., 2016b); and, ultimately, the overall diversity hardly fits into few functional groups (hofmann, 2010; flynn et al., 2012; roselli et al., 2017). plankton play a key role in aquatic ecosystems, being at the base of food-webs and driving biogeochemical cycles, and are experimenting strong perturbations apparently connected to anthropogenic factors, but the fine-scale ecological mechanisms at the base of such phenomena are not fully understood (behrenfeld and boss, 2013; hutchins and fu, 2017; steinberg and landry, 2017). in this context, ‘eco-evo’ approaches, being mainly focused on time (the main dimension of evolution), would be suitable to investigate cause-effect relationships within the wide array of potentially inter-dependent ecological phenomena. long term ecological research (lter), consisting in sampling and analysing physical, chemical and biological variables at fixed sampling sites, with high time-frequency (e.g., weekly), and in the long term (decades), can represent profiting case studies to this readvances in oceanography and limnology, 2017; 8(2): 187-198 review doi: 10.4081/aiol.2017.7194 this work is licensed under a creative commons attribution-noncommercial 4.0 international license (cc by-nc 4.0). biological complexity behind plankton system functioning: synthesis and perspectives from a marine long term ecological research domenico d’alelio* stazione zoologica anton dohrn, villa comunale, 80121 naples, italy *corresponding author: dom.dalelio@gmail.com abstract the functioning of natural communities is the cumulative outcome of multifaceted and intersecting ecological and evolutionary processes occurring at species level. species are not stable entities but evolve in consequence of contingent factors including the relationships they establish with the environment and other co-occurring species. studying ecosystems with an eco-evo approach, i.e., by explicitly considering species evolution and interactions, is thus an essential step to envisioning community adaptation to environmental changes. such an approach would be particularly suitable for studying plankton, a community of both rapidly evolving and strongly interconnected species. in this context, long term ecological research studies (lter) allow investigating nature at different levels of complexity, from species to ecosystems. herein, i examine the most recent results coming from the three-decades plankton lter ‘marechiara’ (lter-mc) in the gulf of naples (mediterranean sea, italy) and discuss their suitability in deepening knowledge on: i) evolutionary bases to plankton diversity (i.e., the founding property of both species and community adaptive potential); ii) ecological and evolutionary determinants of population and community dynamics; and iii) biological complexity behind plankton system functioning. key words: plankton; ecology; evolution; coastal ecosystems; biocomplexity. received: november 2017. accepted: december 2017. no nco mm er cia l u se on ly d. d’alelio188 spect (hughes et al., 2017). by means of lter-based systems ecology analyses, integrating fine-scale biological complexity and biogeochemical processes at ecosystem level, reductionist and holistic approaches congregate, allowing to ‘uncovering the processes hidden because they occur slowly or because effects lag years behind causes’ (magnuson, 1990). this paper takes the three-decades plankton lter ‘marechiara’ (lter-mc) in the gulf of naples (mediterranean sea, italy, fig. 1; ribera d’alcalà et al., 2004) as a benchmark for new-generation lter-based eco-evo studies. this latter approach is far more important in light of: i) the observed fast adaptation of planktonic microbes to global change; ii) the rising impact of the latter on fishery-dependent human societies and iii) the under-exploitation of lter studies in marine policy (barange et al., 2014; irwin et al., 2015; hughes et al., 2017; hutchins and fu, 2017). based on studies carried out in the gulf of naples (gon) and published mostly within the last ten years into isi journals, i herein examine: i) the evolutionary bases to plankton biodiversity (i.e., the founding property of both species and community adaptive potential); ii) the ecological and evolutionary determinants of population and community dynamics; and iii) the suitability of holistic lter-based eco-evo approaches towards understanding the biological mechanisms behind systemic response of plankton to environmental variability. fig. 1. map of the gulf of naples (thyrrenean sea, mediterranean sea, italy) and geographic position of the long term ecological research station marechiara (lter-mc). no nco mm er cia l u se on ly biological complexity behind plankton system functioning 189 evolutionary complexity behind diversity the huge species diversity shown by plankton intrigues ecologists since decades. why does competitive exclusion (i.e., the dominance of few fitter species) does not apply within an unstructured, homogeneous environment like the planktonic one is the main issue introduced by the renowned paper ‘the paradox of the plankton’ by hutchinson (hutchinson, 1961). experimental observations explicated the above-mentioned paradox by suggesting that stable coexistence of distinct species at the same trophic level is a probable outcome of resource-competition processes (tilman 1976, 1977). in this context, more than 500 plankton taxa were identified in the gon within thirty-three years of observation and most of them have apparently similar ecological roles (mazzocchi et al., 2011; piredda et al., 2017). nowadays we now that a high functional redundancy at community level is not unusual in nature (lefcheck et al., 2015), it can emerge even at stable environmental conditions, as reported in experimental evolution studies with microbes (maharjan et al., 2007), and its main ecological role is to guarantee the survival of functional groups in case of species extinctions (rosenfeld, 2002). protists (i.e., unicellular eukaryotes) are the main contributors to plankton metagenome (de vargas et al., 2015; sunagawa et al., 2015). despite the dominance of fastreplicating and bloom-forming species, about 30% of planktonic protist diversity in the oceans is assigned to rare taxa (i.e., whose abundances are less than 0.01% of the total abundance) and the latter can contribute up to 16% of coastal phytoplankton biomass (ignatiades and gotsis-skretas, 2013; de vargas et al., 2015). both blooming and non-blooming species can be present within the same community and at the same time in coastal plankton systems such as the gon (ribera d’alcalà et al., 2004). these data are in line with the observation that a myriad of species in nature are rare (have either low reproductive or high turnover rates) but they are anyway ecologically successful and determinant in community functioning (jain et al., 2014). while the role of the rarest is still not clear in marine plankton, it has been suggested that freshwater ecosystems’ resilience is strongly linked with the presence of rare phytoplankton taxa (downing and leibold, 2010). most evolutionary models based on the classic ‘fitness landscape’ conceptual scheme (wright 1932) and exploiting experimental evolution indicate that microbial species emerge by the divergence of lineages due to differential adaptation to distinct environmental conditions (de visser and krug, 2014). in the above-mentioned model, the fittest clonal lineages are advantaged in respect to ‘flattest’ ones, i.e., those having lower abundance and ecological specialization. such ‘adaptive’ dynamics can be found in data generated by culture-based experimental evolution and genomics involving planktonic protists (lohbeck et al., 2012; mock et al., 2017) and plausibly represents the mechanism behind the fast adaptation of phytoplankton to global change, which is particularly relevant for blooming species (irwin et al., 2015). yet, the above-mentioned ‘fitness model’ apparently does not fit in real frequency-distributions pertaining planktonic protists. some other computational evolutionary-models assess that ‘fittest’ and ‘flattest’ clonal lineages can alternate in dependence of vegetative growth and evolutionary rates (wilke et al., 2001) (fig. 2). namely: i) the fittest emerge at lower mutation and higher replication rates, when rarer positive mutations produce genotype-clouds whose frequency distributes around narrow fitness peaks, while purifying selection sharpens distribution shoulders (wilke et al., 2001); ii) by contrast, the flattest emerge at higher mutation and lower replication rates, when more frequent positive mutations produce genetic clouds including a higher number of slightly different and evenly represented genotypes whose abundances distribute around ‘mutationally robust’ flatter peaks (wilke et al., 2001). in this context, metabarcoding suggests that protist diversity in the gon is higher during winter, i.e., the nonblooming season, when virtually all detected species are rare and blooms are of lower intensity than in other seasons (piredda et al., 2017). by combining the ‘flatness model’ mentioned above with time-repeated biodiversity explorations carried out at lter-mc, one can depict a possible scenario behind the dominance of the flattest during winter non-blooming phases in the gon, a pattern that is explainable with few conceptual steps: • the lower amount of nutrients in the photic zone during winter promotes growth of protist groups with higher surface-to-volume ratios and, thus, higher efficiencies in nutrient-uptake, such as flagellates (zingone et al., 2009; edwards et al., 2013); • while stronger in some characters, selection may be relaxed in other ones, thus leaving room to intra-group genetic divergence and producing flat but ‘rugged’ fitness landscapes not necessarily determined by differential adaptation (koonin, 2011); • in relation with the latter point, despite a common cell shape, (dino)flagellates show high interand intra-specific diversity (gribble and anderson, 2007; murray et al., 2012), which can correspond to potentially functional diversity, such as that present in the production of secondary metabolites (murray et al., 2012). in addition to the simple scenario depicted above, one must consider that even short-term environmental variability strongly contributes in shaping diversity within communities of planktonic protists. for instance, the freno nco mm er cia l u se on ly d. d’alelio190 quent alternations between coastal and offshore waters occurring in the gon during summer, known as ‘green-blue swings’, are a determinant factor for diversity conservation (cianelli et al. 2017; d’alelio et al., 2015): by exerting a so-called ‘intermediate disturbance’ (sensu reynolds et al., 1993), green-blue swings dilute the abundance of opportunistic (most abundant) species and promote survival of the rarest ones (cianelli et al., 2017). in addition, according with integrative studies on microbes coupling experiments and modelling, the fitter-flatter species coexistence emerges within the same environment when copious trade-offs between potentially different strategies establish due to fine-scale environmental variability (beardmore et al., 2011). life-cycle driven evolutionary complexity functional diversity can also evolve by chance (gould, 2002; koonin, 2011). an example to this respect is exaptation, which occurs when phenotypic traits apparently selected for a specific function assume a different and more determinant role in the course of species’ evolutionary history (gould and vrba, 1982). to this respect, adaptation to stochastic environmental factors could be largely dependent from evolvability, or, the capability of a population to generate diversity, enhance the standing genetic variation and develop adaptive solutions (koonin, 2011 and reference therein). evolvability is promoted by the interplay between stochastic biological processes, such as genetic mutation and recombination (koonin, 2011). though conceptually robust, the evolutionary models presented in the previous section do not contemplate homologous genetic recombination, i.e., the exchange of pieces between two similar or identical dna molecules, which constitute an important mechanism of genetic diversification in planktonic prokaryotes and protists (d’alelio and gandolfi, 2012; rengefors et al., 2017). as for planktonic cyanobacteria, genetic mutation and recombination occur at the same rate in the micro-evolution of the freshwater genus planktothrix, but recombination can introduce double more diversity than mutation (d’alelio et al., 2013) and also promote adaptive evolution (tooming-klunderud et al., 2013). concerning protists, many of which have a sexual refig. 2. schematic of evolutionary landscapes (sensu wright 1932) representing evolutionary models potentially applicable to planktonic protists based on observations published by wilke et al. (2001). curves are frequency-distributions for genotypes within different populations. full-grey curves refer to populations at initial conditions (before divergence), empty-black curves refer to populations diverging from those present at initial conditions. no nco mm er cia l u se on ly biological complexity behind plankton system functioning 191 production, recombination occurs with meiosis during gametogenesis (rengefors et al., 2017). in many species, sexual processes are deeply tangled within life-cycles and affect population survival (von dassow and montresor, 2010): this suggests that protist sex is definitely not an evolutionary relict but an acquired strategy with eco-evolutionary implications, both proximate (lineage survival) and ultimate (generating diversity) (speijer et al., 2015). in dinoflagellates, meiosis leads to the production of resting cysts, which are haploid life-cycle stages capable to resist at the sea bottom, so to guarantee survival over longer periods of deprived environmental conditions (von dassow and montresor, 2010). in most diatoms, sex is necessary to generate larger-size cells, thus counteracting the progressive cell-size decrease occurring at each vegetative division and bringing asexual clonal-lineages to death (montresor et al., 2016). the easily-culturable species within the diatom pseudo-nitzschia represent good study-systems in sorting for the role of sex in diatom evolution. for instance, despite the high mutation frequency associated to the dominance of vegetative reproduction (d’alelio et al., 2009a; tesson et al., 2013), periodic sexual events provide species with a cohesive genetic-force that limits intraspecific genetic divergence and promotes species maintenance (amato et al., 2007; d’alelio et al., 2009a). in addition, sex can also occur between different species, leading to hybrid speciation (amato and orsini, 2015; d’alelio and ruggiero, 2015). long-term population genetics of p. multistriata in the gon indicated that planktonic diatoms can produce the same level of diversity (say, genotypic richness) by means of either genetic mutation or recombination (ruggiero et al., in press). in the above-mentioned species, the highest genetic differentiation occurs in the course of apparently infrequent ‘clonal expansions’ establishing when blooms are flanked by a temporary but strong restriction of sex , which determines a positive unbalance of the mutationto-recombination ratio (ruggiero et al., in press) (fig. 3). clonal expansions are ephemeral but massive processes ending with a ‘survival of the fittest’ dynamics that determines a sharp decrease in genotypic richness and the dominance of a ‘super-genotype’, which produces a ‘quasi-monoclonal bloom’ that follows a ‘multi-clonal’ one (ruggiero et al., in press) (fig. 3). similar dynamics are also observed in ‘epidemic’ biofig. 3. schematic of a clonal expansion in a planktonic protist, modified from ruggiero et al. (in press). white lines are genotypes undergoing clonal divergence (lines’ bifurcation), the red line represents a successful genotype emerged in the course of a clonal expansion. reticulated patterns indicate the combined action of genetic divergence (mutation) and convergence (recombination). dotted white and red lines indicate a possible pathway of recombination between white and red lineages after the bloom of the latter. no nco mm er cia l u se on ly d. d’alelio192 logical systems, such as pathogenic bacteria and protozoans, in which clonal expansions are generally linked to strong selection over fitter recombinant genotypes operated by favourable environmental conditions (maynard smith et al., 1993; tibayrenc and ayala, 2012). the clonal expansion detected in p. multistriata was apparently generated by the presence, in the blooming population, of a single mating type, which was therefore unable to undergo sex: the dominance of a ‘super-genotype’ was plausibly determined by life-cycle characteristics as well as by environmental selection (ruggiero et al., in press). the ‘evolutionary jump’ gathered by p. multistriata via clonal expansion made the population dramatically change its genetic fingerprint when the dominant genotype finally recombined with more distantly related ones (ruggiero et al., in press) (fig. 3). microevolutionary bases of population dynamics whereas evolutionary complexity sustains species evolution, diversification processes occurring at population level (i.e., microevolution) promote species adaptation to local conditions. although plankton species’ populations are characterized by a high temporal and spatial intermittency (martin et al., 2005; cloern and jassby, 2010), periodic seasonal blooms are observed at both local and global scales (ruggiero et al., 2015; boyce et al., 2017), suggesting that adaptive processes are at the base of the phenology observed. in confined aquatic systems, such as freshwater lakes, blooms of planktonic cyanobacteria can be reliably linked to evolutionary adaptation (d’alelio et al., 2011): specifically, populations of planktothrix rubescens living in deeper lakes evolved more robust gas-vesicles (i.e., capable to resist stronger water-pressures during lake overturns) than populations living in shallower lakes, and this ‘differential selection’ led distinct populations to float and bloom in the surface photic zone of lakes with different maximum depths. analogous studies have not been performed on planktonic protists, for the lack of reliable molecular resources (i.e., background description of functional loci) that allow tracking differential selection by means of ‘simpler’ population genetics approaches. nonetheless, population genetics focused on neutrally evolving genes, when coupled with life-history investigations, can provide insights into those microevolutionary processes occurring at species level and potentially affecting population dynamics (ruggiero et al., in press). this latter integration can be more likely obtained in lter investigations. biologically regulated life-history processes (such as timing of recruitment of new individuals) are factors reinforcing ecological specialization in general (poisot et al., 2011). among unicellular plankton, diatoms show highly organized life cycles, with a biological clock that regulates the emergence of sex and periodical recruitment of sexual generations (montresor et al., 2016). in the gon, a population of the diatom p. multistriata observed for ten consecutive years underwent sex with a tight biennial periodicity, with two consecutive sexual events separated by about 50 mitotic generations (d’alelio et al., 2010). this biological clock was apparently regulated by cell-size, since sex occurred in cells below a threshold size reached after a precise number of vegetative divisions (d’alelio et al., 2009b). considering planktonic diatoms as model systems for intersecting evolutionary and ecological processes at population level, one may speculate that clonal expansions as that mentioned in the former section can lead to the fast evolution of genotypes particularly adapted to specific environmental conditions, or ecotypes, which can potentially turn into ecological species. based on population biology (genetics and demography) and modelling observations, a possible coupled microevolutionary/life-history dynamics at the base of ecological specialization in the genus pseudo-nitzschia can be drawn as follows: • a cloud of closely related genotypes emerges from a clonal expansion (i.e., a bloom including only closelyrelated genotypes) and constitutes a potential new ecotype (ruggiero et al., in press); • sex occurs at the end of this bloom, when i) encounter between mating cells is favoured by higher population density (d’alelio et al., 2009b) and ii) gametogenesis is energetically affordable because vegetative growth has stopped (scalco et al., 2014); • the sexual progeny enters a precise life-cycle periodicity with sex limited to the blooming season that generated it (d’alelio et al., 2010); • the phasing of bloom and sex promotes breeding and recombination within and not between different ecotypes, thus guaranteeing the maintenance of selected genetic features. since the timing of diatom sex is biologically determined, an ecotype may phase its life cycles with the periodicity of seasonally-determined environmental constraints, thus contributing to the emergence of seasonality. for instance, it has been observed that species in the genus pseudo-nitzschia tend to form blooms during different seasons within the same coastal system (ruggiero et al., 2015) and an incipient ecological speciation, apparently driven by sexual isolation between differently occurring morphotypes, has been observed in p. galaxiae (cerino et al., 2005). the accumulation, generation after generation, of life history processes (such as genetic differentiation, differential adaptation and life-cycle shifts) plausibly provided bases to emergence of seasonality of different plankton no nco mm er cia l u se on ly biological complexity behind plankton system functioning 193 groups in coastal systems like the gon (modigh, 2001; ribera d’alcalà et al., 2004; zingone et al., 2009; mazzocchi et al., 2011; ruggiero et al., 2015; piredda et al., 2017). nonetheless, phenology is subordinated to the ability of species populations to overcome short-term variabilities stemming from the tangled interplay of proximate biological and physical factors whose relative strengths can change seasonally (smayda, 1980; reynolds, 1984; wyatt, 2014). recent advances in disentangling biological from physical drivers of plankton dynamics at short-time scale have been gathered by integrating oceanographic and ecological observations with modelling (cianelli et al. 2017). the application of these techniques to plankton in the coastal gon revealed that the dynamics of species abundances in coastal water masses is mainly ruled by biological factors, such as i) highly plastic physiological-responses of phytoplankton to short-term environmental variability, ii) biologically-regulated germination and formation of resting stages involving (fitter) species producing massive blooms (montresor et al., 2013), and iii) inter-specific interactions, involving mainly non-blooming (flatter) species (cianelli et al. 2017). all these factors ultimately determine species succession at homogeneous environmental conditions (scheffer et al., 2003). from evolutionary to systems ecology ecological communities are complex adaptive entities in which both direct and indirect inter-specific interactions shape the coevolution of complementary traits that promote community stability (joppa et al., 2009; turcotte et al., 2012; guimarães et al., 2017). despite most studies (including those mentioned in the previous sections) consider single species populations in a simplified context ruled by genetic and environmental constraints, plankton species are not mutually isolated in the environment, their populations are continuously mixed one another and several kinds of interactions can establish and affect both population and community dynamics (lima-mendez et al., 2015; d’alelio et al., 2016a). coevolution of plankton organisms has been mainly put in relation to mutualistic and antagonistic interactions, such as symbiosis and parasitism, which seem to be widespread in the oceans and play an important role in global biogeochemical cycles (lima-mendez et al., 2015; guidi et al., 2016). nonetheless, other trophic relationships can emerge from complex natural history processes in which predators and preys reciprocally affect each other’s evolution. for instance, the pelagic tunicate oikopleura dioica (appendicularia) is capable of ‘breeding’ the ciliate strombidium spp. with plankton particles directed by feeding currents towards the tunicate’s gelatinous ‘house’ (lombard et al., 2010); therefore, when its esophagus has grown enough to ingest larger particles, the same appendicularian feeds on ciliate cells, which are energetically richer than small-sized phytoplankton (lombard et al., 2010). remarkably, parasitism, mutualism and predatorprey relationship succeed in time in the course of a single life-history (lombard et al., 2010). ciliate-appendicularian coevolution apparently emerged from the mutual ecological benefit of establishing a trophic interaction: namely, ciliates provide appendicularians with an essential, additional food-supply and, at the same time, take advantage of appendicularian houses to survive and grow in food-limited environments (lombard et al., 2010). furthermore, this evolutionarydetermined ecological strategy has important implications in the functioning of the pelagic system, since appendicularians and ciliates play an important position in plankton food-web (d’alelio et al., 2016a). while appendicularians are important hubs (i.e., they up-take and deliver a remarkable amount of organic matter within the plankton food-web) and act as keystone species (sensu power et al., 1996), ciliates are important food-sources for copepods in oligotrophic conditions, because they deliver to these latter animals organic matter which they cannot directly eat, such as the smaller-sized picoplankton (d’alelio et al., 2016a). interaction-based plankton functioning plankton are considered as ‘complex adaptive systems’ (leibold and norberg, 2004), in which low-level interactions, i.e., between individuals and the environment and among individuals, determine high-level collective responses (levin, 2007). therefore, investigating shortterm system-responses can provide us with conceptual bases to delineate the self-organization/regulation abilities of plankton communities. to the latter respect, the coastal plankton in the gon, which lives within an instable environment at the boundary between the coastal, eutrophic, and offshore, oligotrophic, dominions, can represent a suitable model system (ribera d’alcalà et al., 2004; d’alelio et al., 2015). in the gon, a study comparing community dynamics from nanoflagellate to predatory mesozooplankton (within an individual size-range spanning 5-2·103 µm) indicated the presence of co-variations of species trends potentially related to different trophic links (d’alelio et al., 2015). when assembled into networks, co-variation links help identifying system responses to a level higher than that of population dynamics (loreau, 2010). for instance, the association network referring to a seasonal plankton community in the gon displayed a vertical topology (i.e., phytoplankton => herbivore zooplankton => carnivore no nco mm er cia l u se on ly d. d’alelio194 zooplankton) during coastal, eutrophic states and a more scattered topology (dominated by links among microbes) during offshore, oligotrophic states (d’alelio et al., 2015). the above-mentioned study suggested that: • the effects of resource intermittency may propagate at different levels of ecological complexity (from individuals to individuals’ interactions); • the community may respond ‘adaptively’ to physicalchemical changes, like oscillations between euand oligotrophy; • biological diversity would be crucial to guarantee a system-response driven by changes in trophic interactions within the same community. biodiversity is thought to enhance trophic diversity within ecosystems (lefcheck et al., 2015), and, in turn, the stability of the latter depends from the robustness of ecological networks regulated by trophic interactions (barabás et al., 2017). studying food-webs is thus fundamental to reconcile the biodiversity (how many species are there), structure (how they relate reciprocally) and functioning (what they do collectively) in ecosystems (thompson et al., 2012). also, ecological network models exploiting the ecopath methodology (i.e., interpolating ecological networks by modelling biomass flows into ecosystems) are among the best tools to this purpose (barabás et al., 2017). building on qualitative observations carried out in the gon, the plankton community was investigated with one of such models (d’alelio et al., 2015, 2016a, 2016b), in which: • ‘diversity’ was defined by the variety of ‘functional web-nodes’, i.e., species or groups of organisms with specific biological characteristics (namely, size, physiology, metabolism, behaviour and diet); • ‘structure’ was defined by the topology of food-weblinks, i.e., overall direction and intensity of biomass fluxes among web-nodes (derived iteratively by the model based on nodes’ biomass and biological characteristics); • ‘functioning’ was defined as the efficiency of fluxes (across the web and between consecutive trophic levels) that was estimated from model output. the above-mentioned model reproduced a plankton food-web including very few specialists, which limited interspecific competition, and a huge amount of weak trophic links, which increased trophic alternatives (fig. 4). in virtue of these properties, almost all species in the plankton food-web could switch their trophic preferences based on available resources (fig. 4). the above-mentioned study indicated that, when integrated within a foodweb context, evolutionary determined ecological strategies were crucial to drive system functionality (d’alelio et al., 2016a): • firstly, nested and convoluted protozoan-metazoan interactions involved a myriad of trophic strategies (such as mixotrophy, niche partitioning among protozooplankters and different selective feeding by mesozooplankton) establishing several potential trophic pathways; • the above-mentioned trophic step showed the highest trophic efficiency (up to 25%), the latter being the ratio between the biomass taken by a trophic level and that delivered to the subsequent one; • high efficiency at intermediate trophic steps allowed smoothing the effects of oscillations in primary production on planktivorous-fish production; • finally, the surplus of matter and energy available at the lower levels of the web was used by protozooplankton as a resource to maintain species diversity. trophic plasticity at organismal level determines ecological network flexibility, expressed as modifications in both direction and intensity of trophic links (fig. 4): this allows plankton food-web to respond adaptively to system changes. such mechanism can explain the high resilience of mesozooplankton grazers to trophic intermittency reported in the gon (mazzocchi et al., 2012) (d’alelio et al., 2016a). concluding remarks important modifications are presently occurring in marine plankton communities e.g., the rise of harmful algal blooms and global decrease of both phytoand zooplankton biomasses (boyce et al., 2010; chust et al., 2014; glibert and burford, 2017). these phenomena are plausibly the result of complex feedback mechanisms determined by interplaying biological and environmental factors, such as fast adaptation of microbes to changes in chemical-resources regimes and trophic cascades occurring at food-web level. despite the rising of reductionist evolutionary-ecology approaches (mainly focusing on phytoplankton experimental evolution and population dynamics; e.g., collins et al., 2014 and references therein), holistic ‘systems ecology’ approaches that explicitly consider species interactions in the process of understanding of plankton functioning and resilience to environmental changes are still at their infancy (stec et al., 2017). indeed, studying ecosystems from a time-based, evolutionary perspective relies on the availability of data over a time-period that is suitable to observing processes that occur at very different time scales but are all interconnected. in fig. 5 the main cause-effects relationships playing in plankton function discussed in the present paper have been assembled. being based on patterns observed and processes identified in the gon, some links in the abovementioned network were not discussed though, such as the effects of genetic diversity on reproduction rates and of environmental factors on population dynamics and no nco mm er cia l u se on ly biological complexity behind plankton system functioning 195 fig. 4. schematic of the plankton food-web in the gulf of naples during oligotrophic and eutrophic states, modified from d’alelio et al. (2016a). nodes are species or group of species; links are biomass fluxes. in order to enhance data visualization, web links have been obtained from log-transformation of fluxes-data presented in d’alelio et al. (2016a). fig. 5. hierarchical network schematizing the main regulative mechanisms of plankton system functioning based on considerations presented in this paper. full and dotted arrows illustrate direct and indirect links, respectively. the dotted square includes community properties indirectly affected by community functioning. no nco mm er cia l u se on ly d. d’alelio196 inter-specific interactions (for a review, see d’alelio et al., 2016b and references therein). one must anyway consider that biocomplexity in the plankton system largely exceeds that considered in the present paper. as for evolutionary complexity and its role in species adaptation, microevolutionary models discussed herein can be likely applied to asexually reproducing planktonic animals (e.g., cladocerans and pelagic tunicates) in the need of interpreting their adaptation dynamics to changing environmental conditions. concerning ecological complexity, plankton food-web models should i) consider presently neglected organisms, such as virus and jellyfish, both playing fundamental roles in marine ecosystems (boero, 2015; lara et al., 2017), and ii) integrate also the benthos dominion, in light of frequent biologically-mediated interactions between the two systems (d’alelio et al., 2017). to this latter respect, lters offer a unique opportunity for investigating ecological and evolutionary determinants driving plankton functioning, from reproductive processes occurring at species level to the circulation of energy and matter playing at system level (fig. 5). these researches can set important conceptual and methodological backgrounds to next-generation observatory studies exploiting meta-omics technologies (de vargas et al., 2015; sunagawa et al., 2015; guidi et al., 2016) and exploring plankton biocomplexity and functioning over complete and time-resolved biological and physical data matrices. acknowledgments this work is dedicated to the memory of dr. giuseppe morabito, a brilliant plankton ecologist as well as a great man. one anonymous reviewer is gratefully acknowledged for providing comments improving the quality of this work. the lter-mc belongs to the national and international long term ecological research networks (lter-italy, lter-europe and ilter). lter-mc data generation and processing is entirely supported by the stazione zoologica anton dohrn, naples. i thank the flagship project ritmare the italian research for the sea coordinated by the italian national research council and funded by the italian ministry of education, university and research within the national research program 2011-2013. references amato a, kooistra whcf, ghiron jhl, mann dg, pröschold t, montresor m, 2007. reproductive isolation among sympatric cryptic species in marine diatoms. protist 158:193-207. amato a, orsini l, 2015. rare interspecific breeding in pseudo-nitzschia (bacillariophyceae). phytotaxa 217:145. barabás g, michalska-smith mj, allesina s, 2017. self-regulation and the stability of large ecological networks. nat. ecol. evol. barange m, merino g, blanchard jl, scholtens j, harle j, allison eh, et al., 2014. impacts of climate change on marine ecosystem production in societies dependent on fisheries. nat. clim. chang. 4:211-216. beardmore re, gudelj i, lipson da, hurst ld, 2011. metabolic trade-offs and the maintenance of the fittest and the flattest. nature 472:342-346. behrenfeld mj, boss es, 2013. resurrecting the ecological underpinnings of ocean plankton blooms. ann. rev. mar. sci. 1-28. boero f, 2015. the future of the mediterranean sea ecosystem: towards a different tomorrow. rend. lincei 26:3-12. bone q, others, 1998. the biology of pelagic tunicates. oxford university press oxford. boyce dg, frank kt, leggett wc, 2015. from mice to elephants: overturning the “one size fits all” paradigm in marine plankton food chains. ecol. lett. 18:504-515. boyce dg, lewis mr, worm b, 2010. global phytoplankton decline over the past century. nature 466:591-596. boyce dg, petrie b, frank kt, worm b, leggett wc, 2017. environmental structuring of marine plankton phenology. nat. ecol. evol. 1:1484. cerino f, orsini l, sarno d, dell’aversano c, tartaglione l, zingone a, 2005. the alternation of different morphotypes in the seasonal cycle of the toxic diatom pseudo-nitzschia galaxiae. harmful algae 4:33-48. chust g, allen ji, bopp l, schrum c, holt j, tsiaras k, et al., 2014. biomass changes and trophic amplification of plankton in a warmer ocean. glob. chang. biol. 20:2124-2139. cianelli d, d’alelio d, uttieri m, sarno d, zingone a, zambianchi e, ribera d’alcalà m, 2017 disentangling physical and biological drivers of phytoplankton dynamics in a coastal system. sci. rep. 7:1-15 cloern je, jassby ad, 2010. patterns and scales of phytoplankton variability in estuarine-coastal ecosystems. estuaries and coasts 33:230-241. collins s, rost b, rynearson ta, 2014. evolutionary potential of marine phytoplankton under ocean acidification. evol. appl. 7:140-155. d’alelio d, amato a, kooistra whcf, procaccini g, casotti r, montresor m, 2009a. internal transcribed spacer polymorphism in pseudo-nitzschia multistriata (bacillariophyceae) in the gulf of naples: recent divergence or intraspecific hybridization? protist 160:9-20. d’alelio d, amato a, luedeking a, montresor m, 2009b. sexual and vegetative phases in the planktonic diatom pseudo-nitzschia multistriata. harmful algae 8:225-232. d’alelio d, gandolfi a, 2012. recombination signals in the rpoc1 gene indicate gene-flow between planktothrix (cyanoprokaryota) species. j. phycol. 48:1424-1432. d’alelio d, gandolfi a, boscaini a, flaim g, tolotti m, salmaso n, 2011. planktothrix populations in subalpine lakes: selection for strains with strong gas vesicles as a function of lake depth, morphometry and circulation. freshwater biol. 56:1481-1493. d’alelio d, libralato s, wyatt t, ribera d’alcalà m, 2016a. ecological-network models link diversity, structure and function in the plankton food-web. sci. rep. 6:21806. d’alelio d, luongo g, di capua i, 2017. plankton food for benthic fish: de visu evidence of trophic interaction between rainbow wrasse (coris julis) and pelagic tunicates (pegea confoederata). adv. oceanogr. limnol. 8:6973. no nco mm er cia l u se on ly biological complexity behind plankton system functioning 197 d’alelio d, mazzocchi mg, montresor m, sarno d, zingone a, margiotta f, et al., 2015. the green blue swing: plasticity of plankton food-webs in response to coastal oceanographic dynamics. mar. ecol. 36:1155-1170. d’alelio d, montresor m, mazzocchi mg, margiotta f, sarno d, ribera d’alcalà m, 2016b. plankton food-webs: to what extent can they be simplified? adv. oceanogr. limnol. 7:67-92. d’alelio d, ribera d’alcalà m, dubroca l, sarno d, zingone a, montresor m, 2010. the time for sex: a biennial life cycle in a marine planktonic diatom. limnol. oceanogr. 55:106114. d’alelio d, ruggiero mv, 2015. interspecific plastidial recombination in the diatom genus pseudo-nitzschia. j. phycol. 51:1024-1028. d’alelio d, salmaso n, gandolfi a, 2013. frequent recombination shapes the epidemic population structure of planktothrix (cyanoprokaryota) in italian subalpine lakes. j. phycol. 49:1107-1117. de vargas c, audic s, henry n, decelle j, mahé f, logares r, et al., 2015. eukaryotic plankton diversity in the sunlit ocean. science 348:1-12. de visser jag, krug j, 2014. empirical fitness landscapes and the predictability of evolution. nat. rev. genet. 15:480-490. downing amyl, leibold ma, 2010. species richness facilitates ecosystem resilience in aquatic food webs. freshw. biol. 55:2123-2137. edwards kf, litchman e, klausmeier ca, 2013. functional traits explain phytoplankton community structure and seasonal dynamics in a marine ecosystem. ecol. lett. 16:56-63. flynn kj, stoecker dk, mitra a, raven ja, glibert pm, hansen pj, graneli e, burkholder jm, 2012. misuse of the phytoplankton-zooplankton dichotomy: the need to assign organisms as mixotrophs within plankton functional types. j. plankton res. 35:3-11. glibert pm, burford ma, 2017. globally changing nutrient loads and harmful algal blooms: recent advances, new paradigms, and continuing challenges. oceanography 30:58-69. gould sj, 2002. the structure of evolutionary theory. harvard university press, harvard:1464 pp. gould sj, vrba es, 1982. exaptation-a missing term in the science of form. paleobiology 8:4-15. gribble ke, anderson dm, 2007. high intraindividual, intraspecific, and interspecific variability in large-subunit ribosomal dna in the heterotrophic dinoflagellates protoperidinium, diplopsalis, and preperidinium (dinophyceae). phycologia 46:315-324. guidi l, chaffron s, bittner l, eveillard d, larhlimi a, roux s, et al., 2016. plankton networks driving carbon export in the oligotrophic ocean. nature 532:465-470 guimarães pr, pires mm, jordano p, bascompte j, thompson jn, 2017. indirect effects drive coevolution in mutualistic networks. nature 550:511-514. hendry ap, 2016. eco-evolutionary dynamics. princeton university press: 416 pp. hofmann ee, 2010. plankton functional group models an assessment. prog. oceanogr. 84:16-19. hughes bb, beas-luna r, barner ak, brewitt k, brumbaugh dr, cerny-chipman eb, et al., 2017. long-term studies contribute disproportionately to ecology and policy. bioscience 67:271-281. hutchins da, fu f, 2017. microorganisms and ocean global change. nat. microbiol. 2: 201758. hutchinson g. e., 1961. the paradox of the plankton. am. nat. 95:137-145. ignatiades l, gotsis-skretas o, 2013. the contribution of rare species to coastal phytoplankton assemblages. mar. ecol. 35:132-145. irwin aj, finkel z v, müller-karger fe, troccoli ghinaglia l, 2015. phytoplankton adapt to changing ocean environments. p. natl. acad. sci. usa 1-5. jain m, flynn dfb, prager cm, hart gm, devan cm, ahrestani fs, et al., 2014. the importance of rare species: a trait-based assessment of rare species contributions to functional diversity and possible ecosystem function in tallgrass prairies. ecol. evol. 4:104-112. joppa ln, bascompte j, montoya jm, solé r v, sanderson j, pimm sl, 2009. reciprocal specialization in ecological networks. ecol. lett. 12:961-9. koonin e v, 2011. the logic of chance: the nature and origin of biological evolution. ft press: 528 pp. lara e, vaqué d, sà el, boras ja, gomes a, borrull e, et al., 2017. unveiling the role and life strategies of viruses from the surface to the dark ocean. sci. adv. 3:e1602565. lefcheck js, byrnes jek, isbell f, gamfeldt l, griffin jn, eisenhauer n, et al., 2015. biodiversity enhances ecosystem multifunctionality across trophic levels and habitats. nat. commun. 6:1-7. leibold ma, norberg j, 2004. biodiversity in metacommunities: plankton as complex adaptive systems? limnol. oceanogr. 49:1278-1289. levin s, 2007. fragile dominion. basic books. lima-mendez g, faust k, henry n, decelle j, colin s, carcillo f, et al., 2015. determinants of community structure in the global plankton interactome. science 348:1-10. lohbeck kt, riebesell u, reusch tbh, 2012. adaptive evolution of a key phytoplankton species to ocean acidification. nat. geosci. 5:346-351. lombard f, eloire d, gobet a, stemmann l, dolan jr, sciandra a, gorsky g, 2010. experimental and modeling evidence of appendicularian-ciliate interactions. limnol. oceanogr. 55:77-90. loreau m, 2010. from polutations to ecosystems: theoretical fondations for a new ecological synthesis. princeton university press: 320 pp. magnuson jj, 1990. long-term ecological research and the invisible present. bioscience 40:495-501. maharjan rp, seeto s, ferenci t, 2007. divergence and redundancy of transport and metabolic rate-yield strategies in a single escherichia coli population. j. bacteriol. 189:2350-2358. martin ap, zubkov m v, burkill ph, holland rj, 2005. extreme spatial variability in marine picoplankton and its consequences for interpreting eulerian time-series. biol. lett. 1:366-369. maynard smith j, smith nh, o’rourke m, spratt bg, 1993. how clonal are bacteria? p. natl. acad. sci. usa 90:4384-4388. mazzocchi mg, dubroca l, garcía-comas c, capua i di, ribera d’alcalà m, 2012. stability and resilience in coastal copepod assemblages: the case of the mediterranean longterm ecological research at station mc (lter-mc). prog. oceanogr. 97-100:135-151. mazzocchi mg, licandro p, dubroca l, capua i di, saggiomo no nco mm er cia l u se on ly d. d’alelio198 v, 2011. zooplankton associations in a mediterranean longterm time-series. j. plankton res. 33:1163-1181. mock t, otillar rp, strauss j, mcmullan m, paajanen p, schmutz j, et al., 2017. evolutionary genomics of the cold-adapted diatom fragilariopsis cylindrus. nature 541:536-540. modigh m, 2001. seasonal variations of photosynthetic ciliates at a mediterranean coastal site. aquat. microb. ecol. 23:163-175. montresor m, di prisco c, sarno d, margiotta f, zingone a, 2013. diversity and germination patterns of diatom resting stages at a coastal mediterranean site. mar. ecol. prog. ser. 484:79-95. montresor m, vitale l, d’alelio d, ferrante mi, 2016. sex in marine planktonic diatoms: insights and challenges. perspect. phycol. 3:61-75. morabito g, oggioni a, caravati e, panzani p, 2007. seasonal morphological plasticity of phytoplankton in lago maggiore (n. italy). hydrobiologia 578:47-57. murray sa, garby t, hoppenrath m, neilan ba, 2012. genetic diversity, morphological uniformity and polyketide production in dinoflagellates (amphidinium, dinoflagellata). plos one 7:1-14. piredda r, tomasino mp, d’erchia am, manzari c, pesole g, montresor m, et al., 2017. diversity and temporal patterns of planktonic protist assemblages at a mediterranean long term ecological research site. fems microbiol. ecol. 93:1-14. poisot t, bever jd, nemri a, thrall ph, hochberg me, 2011. a conceptual framework for the evolution of ecological specialisation. ecol. lett. 14:841-851. power me, tilman d, estes ja, menge ba, bond wj, 1996. challenges in the quest for keystones. bioscience 46:609. rengefors k, kremp a, reusch tbh, wood am, 2017. genetic diversity and evolution in eukaryotic phytoplankton: revelations from population genetic studies. j. plankton res. 0:1-15. reynolds cs, 1984. phytoplankton periodicity: the interactions of form, function and environmental variability. freshw. biol. 14:111-142. reynolds cs, padisák j, sommer u, 1993. intermediate disturbance in the ecology of phytoplankton and the maintenance of species diversity: a synthesis, p. 183-188 in: j. padisák, cs reynolds and u. sommer (eds.), intermediate disturbance hypothesis in phytoplankton ecology. springer. ribera d’alcalà m, conversano f, corato f, licandro p, mangoni o, marino d, et al., 2004. seasonal patterns in plankton communities in a pluriannual time series at a coastal mediterranean site (gulf of naples): an attempt to discern recurrences and trends. sci. mar. 68:65-83. roselli l, litchman e, stanca e, cozzoli f, basset a, 2017. individual trait variation in phytoplankton communities across multiple spatial scales. j. plankton res. 39:577-588. rosenfeld js, 2002. functional redundancy in ecology and conservation. oikos 98:156-162. ruggiero mv, d’alelio d, ferrante mi, santoro m, vitale l, procaccini g, montresor m, 2017. clonal expansion behind a marine diatom bloom. isme j. doi: 10.1038/ismej. 2017.181. [epub ahead of print]. ruggiero mv, sarno d, barra l, kooistra whcf, montresor m, zingone a, 2015. diversity and temporal pattern of pseudo-nitzschia species (bacillariophyceae) through the molecular lens. harmful algae 42:15-24. scalco e, stec k, iudicone d, ferrante mi, montresor m, 2014. the dynamics of sexual phase in the marine diatom pseudo-nitzschia multistriata (bacillariophyceae). j. phycol. 50:817-828. scheffer m, rinaldi s, huisman j, weissing fj, 2003. why plankton communities have no equilibrium: solutions to the paradox. hydrobiologia 491:9-18. seckbach j, kociolek p, 2011. the diatom world. springer science, berlin: 534 pp. smayda tj, 1980. phytoplankton species succession, p. 493570. in: i. morris (ed.), the physiological ecology of phytoplankton. university of california press, berkeley speijer d, lukeš j, eliáš m, 2015. sex is a ubiquitous, ancient, and inherent attribute of eukaryotic life. p. natl. acad. sci. usa 112:8827-34. stec kf, caputi l, buttigieg pl, d’alelio d, ibarbalz fm, sullivan mb, et al., 2017. modelling plankton ecosystems in the meta-omics era. are we ready? mar. genomics 32:117. steinberg dk, landry mr, 2017. zooplankton and the ocean carbon cycle. ann. rev. mar. sci. 9:413-444. stoecker dk, hansen pj, caron da, mitra a, 2017. mixotrophy in the marine plankton. ann. rev. mar. sci. 9:311-335. sunagawa s, coelho lp, chaffron s, kultima jr, labadie k, salazar g, et al., 2015. structure and function of the global ocean microbiome. science 348:1-10. tesson svm, legrand c, oosterhout c van, montresor m, kooistra whcf, procaccini g, 2013. mendelian inheritance pattern and high mutation rates of microsatellite alleles in the diatom pseudo-nitzschia multistriata. protist 164:89-100. thompson rm, brose u, dunne ja, hall ro, hladyz s, kitching rl et al., 2012. food webs: reconciling the structure and function of biodiversity. trends ecol. evol. 27:689-97. tibayrenc m, ayala fj, 2012. reproductive clonality of pathogens: a perspective on pathogenic viruses, bacteria, fungi, and parasitic protozoa. p. natl. acad. sci. usa 109:e3305-13. tilman d, 1976. ecological competition between algae: experimental confirmation of resource-based competition theory. science 192:463-465. tilman d, 1977. resource competition between plankton algae: an experimental and theoretical approach. ecology 58:338348. tooming-klunderud a, sogge h, rounge tb, nederbragt aj, lagesen k, glöckner g, et al., 2013. from green to red: horizontal gene transfer of the phycoerythrin gene cluster between planktothrix strains. appl. environ. microbiol. 79:6803-6812. turcotte mm, corrin msc, johnson mtj, 2012. adaptive evolution in ecological communities. plos biol. 10:e1001332. von dassow p, montresor m, 2010. unveiling the mysteries of phytoplankton life cycles: patterns and opportunities behind complexity. j. plankton res. 33:3-12. wilke co, wang jl, ofria c, lenski re, adami c, 2001. evolution of digital organisms at high mutation rates leads to survival of the flattest. nature 412:331-333. wright s, 1932. the roles of mutation, inbreeding, crossbreeding and selection in evolution. proceedings 6th ann. congr. of genetics 1:356-366. wyatt t, 2014. margalef’s mandala and phytoplankton bloom strategies. deep sea res. part ii top. stud. oceanogr. 101:32-49. zingone a, dubroca l, iudicone d, margiotta f, corato f, ribera d’alcalà m, et al., 2009. coastal phytoplankton do not rest in winter. estuaries coasts 33:342-361. no nco mm er cia l u se on ly layout 1 introduction annually laminated (varved) lacustrine deposits enable high resolution reconstruction of past environmental and climate changes. since such sediments are relatively rarely deposited in lakes, and are therefore seldom analysed, they represent important material for paleolimnological studies. a number of studies primarily deal with issues related to laminae, such as the process of their formation and preservation (brauer and casanova, 2001; zolitschka et al., 2015; kemp, 2016) and methods for searching for lacustrine sediments, especially those with well-preserved annual lamination (brauer et al., 1999; brauer, 2004; mingram et al., 2007; tylman et al., 2012; wulf et al., 2016). an important issue regards also type and structure of laminae (sturm, 1979; sturm and lotter, 1995; zolitschka, 2007). lake gościąż (poland), which was investigated in the 1990s and presents annually laminated lacustrine sediments deposited during the last ca. 13000 years, has become a reference site for central europe and a stimulus for a number of further extensive search for lakes with visible, well-preserved annually laminated sediments (brauer and casanova, 2001; kinder et al., 2013; tylman et al., 2013). especially lakes with a complete sequence of lamination from their initial period to present day represent the most valuable study material. these deposits allow the annual resolution of the lake chemical and biological (e.g., deposited plant and animal remains) composition, which are recorded by the varves. an important aspect of varved sediments is their seasonal lamination (summer-winter), which implies that both ecological and climatic changes occurring in the past can be followed with an extremely high accuracy (ralska-jasiewiczowa, et al., 1998; szeroczyńska, 1998a; last and smol, 2001; nykӓnen et al., 2010). the remains of invertebrates, cladocera in particular, are an important, well preserved and taxanomically well known autochthonous element of lacustrine sediments, and due to their role of primary consumers they are considered to be important bioindicators of both bottom up and top-down ecological drivers of lakes (hann et al., 1994; manca and comoli, 1996; szeroczyńska, 1998b; jeppesen et al., 2000; korhola and rautio, 2001; chen et al., 2010; niska and mirosławgrabowska, 2015). sediment remains of cladocera are one of the basic elements of the paleolimnological analysis (boucherle and züllig, 1983; hall and smol, 1996; van damme and kotov, 2016). advances in oceanography and limnology, 2016; 7(2): 184-196 article doi: 10.4081/aiol.2016.6297 this work is licensed under a creative commons attribution-noncommercial 4.0 international license (cc by-nc 4.0). long term subfossil cladocera record from the partly varved sediment of lake tiefer see (ne germany) krystyna szeroczyńska* institute of geological sciences, polish academy of sciences, research centre warsaw, twarda 51/55, pl 00818 warsaw, poland *corresponding author: kszerocz@twarda.pan.pl abstract the partly varved and well-dated sediment record of lake tiefer see (ne germany) allowed the high resolution paleolimnological reconstruction of the lake evolution during the whole holocene. this paper presents results of subfossil cladocera analysis. during the holocene, the fauna of subfossil cladoceran was represented by 36 species belonging to 6 families. cladocera were dominated by typical open-water species, belonging especially of the bosminidae family. the sediment record of lake tiefer see exhibited distinct decadalto centennial-scale alternations of welland non-varved intervals, which were related to changes in the thermal circulation of the lake water column. in general, well varved sediments were deposited during periods of reduced lake circulation, and were characterised by maximum abundance of cladocera, whereas non-varved sedimentation phases occurred during periods of increased lake circulation and showed a lower number of cladocera specimens. the most suitable conditions for the development of cladoceran fauna occurred during the early holocene and from ~2055 – 725 yr cal bp. on the basis of the increasing number of species associated with high lake productivity, eight stages of increasing trophy were inferred. the first two were attributed to climate warming, while the next six to human impact. higher human driven trophic conditions of lake tiefer see occurred in the periods 5750-5500 and 4500-4100 yr cal bp, and four times from 1000 to 50 yr cal bp. during the last 750 years and in the period from 6000 to 2500 yr cal bp, the species of eubosmina produced extreme morphs. the cyclomorphosis of eubosmina was likely connected to more pronounced changes in the lake environmental conditions key words: lake tiefer see; holocene; subfossil cladocera; eubosmina morphs. received: september 2016. accepted: november 2016. no n c om me rci al us e o nly long term subfossil cladocera from lake tiefer see 185 the analysis of subfossil cladocera was performed for a number of lake locations in europe, but mostly nonvarved deposits were analysed (sarmaja-korjonen, 2001; kamenik et al., 2007; manca et al., 2007; zawisza and szeroczyńska, 2007; bennion et al., 2011; kulesza et al., 2011; mirosław-grabowska and zawisza, 2014; zawiska et al., 2015; nevalainen and luoto, 2016; milan, 2016). on the contrary, studies on the subfossil cladocera fauna in varved sediments was conducted for only some locations in europe so far, mainly in germany, poland, switzerland and scandinavia (hofmann, 1993a, 2001; szeroczyńska, 1998a; nykӓnen et al., 2010). the annually laminated sediment record from lake tiefer see allowed the detailed reconstruction of climate and environmental changes during the holocene (kienel et al., 2013; drӓger et al., 2016; wulf et al., 2016). the present study presents the results of subfossil cladocera analysis performed on the entire sediment record from lake tiefer see (ne germany). the main objectives of this study were to reconstruct the long term evolution of the cladocera community and to describe the ecological conditions of the lake from its origin to the modern time. furthermore, this study aimed also at comparing the results obtained from the analysis of cladocera remains collected from lake tiefer see, with those previously obtained from lake gościąż (central poland). lake gościąż was selected for its comparable geographical location and environmental features respect to lake tiefer see and for the fact that so far this represents the sole lakes with almost entirely varved sediments, where analysis of subfossil cladoceran was performed. methods principal study sites: lake tiefer see lake tiefer see is located in ne germany (53°35.5’ n, 12°31.8’ e) in the region of mecklenburg vorpommern (fig. 1). the lake originated during the last glaciations and well-varved, poorlyvarved and non-varved sediments fig. 1. a) location of lake tiefer see (germany) and lake gościąż (poland). b) bathymetric map of lake tiefer see. c) bathymetric map of lake gościąż. no n c om me rci al us e o nly 186 k. szeroczyńska were deposited in the lake since its initial stages. lake tiefer see is located at an altitude of 62 m asl, has an area of about 0.75 km2, and a maximum depth of 62 m. the lake is 1 km long and its maximum width is around 400 m. the lake is located in the german region affected by oceanic climate, where average january air temperature is 0°c and july temperature about 18°c. currently the lake is a dimictic to monomictic, mesotrophic, with electric conductivity of 575 µs cm–1 (kienel et al., 2013; drӓger et al., 2016). in 2011 and 2013 several sediment cores were collected, on which basis a composite 1083 cm long profile was constructed (dräger et al., 2016). due to sediment loss during coring, the sediment profile contains two gaps (at 769.5 cm 956.5 cm depth), each one probably of several decimetres length (fig. 2). due to these gaps a continuous composite profile could only be constructed for the upper 770 cm of the sediment profile, covering the past ca. 6000 years. the chronology for the upper part of the profile was established by a multiple dating approach, including varve counts, ams 14c dating, and tephrochronology (kienel et al., 2013; drӓger et al., 2016; wulf et al., 2016). chronology uncertainties vary along the sediment record and amount to ±85 years at the base of the studied interval at 6030 yr cal. bp. the poor age control in the lower part of the sediment profile, i.e. below 770 cm depth only, allowed a descriptive data evaluation. the lacustrine sediment deposition started at about 1070 cm depth, and early sediments were mainly composed of calcite, plant remains and minerogenic detritus. from ca.1040 cm depth upwards, organic rich sediments were characterized by alternating well-varved, poorly varved and non-varved sediment sequences. well-varved sediments are mainly composed of sublayers formed by carbonates (calcite and ca-rhodochrosite), diatoms and organic matter (drӓger et al., 2016; wulf et al., 2016). poorlyand non-varved sediment sections are enriched in quartz grains, plant fragments, benthic diatoms and occasionally bivalves. all ages in this study are given as calibrated years before present (yr cal bp), and as anno domini (ad) years in the historic period. site for comparison study: lake gościąż lake gościąż (52°35’ n, 19°21’ e) is located on the vistula terrace (central poland, fig. 1) and belongs to a complex of four connected lakes (na jazach lake system, gostynińskie lake district). the lake is located at 64.3 m asl, its current area is 41.7 ha, its maximum depth is 24 m, and its mean width and length are ~400 m and 1168 km, respectively. the lake is located in the oceanic climate zone with average january air temperature of -2.8°c fig. 2. percentage composition of the subfossil cladocera species, and total cladocera sum in the sediments of lake tiefer see (composite profile). dotted vertical grey arrows (on bosmina longirostris) indicated stages of trophic increase. 1, section with well developed varved (regular laminae); 2, non-varved section; 3, poorly varved section; 4, 11,492±253 yr cal bp±2σ (drӓger et al., 2016). no n c om me rci al us e o nly long term subfossil cladocera from lake tiefer see 187 and july temperature of ca. 18°c (ralska-jasiewiczowa et al., 1998), and is currently in mesotrophic to eutrophic conditions. lake gościąż was formed during the last glaciations. during the late glacial and the holocene, sediments of gyttia type were deposited in the lake. they are well varved, except for the youngest sediments (especially since about 1550 yr cal bp, profile g1/87). cladocera analysis the analysis of subfossil cladocera was performed at 2-5 cm resolution for the last 6000 years of sedimentation at lake tiefer see, i.e. from 0 to 770 cm sediment depth, while the sediments below 770 cm depth were analysed at 5-10 cm resolution (700-1065 cm). due to the two gaps located at ca. 770 and 957 cm depth (fig. 2), a robust chronology exists only for the upper 770 cm of the sediment profile (drӓger et al., 2016). therefore, the cladocera data obtained from the topmost 770 cm are presented both with respect to the time and depth scale, whereas data below 770 cm are presented only along the depth scale, and hence are only presented in a descriptive way. sediments were analysed for physical parameters and mineralogical composition (drӓger et al., 2016). the analysis of subfossil cladocera was performed in accordance with the generally accepted standard methods (frey, 1986; korhola and rautio, 2001). after removal of carbonates, each sample (1 cm3 of fresh sediment) was macerated in a 10% koh solution, and washed through a 33 µm sieve. the obtained residuum was analysed under an optical microscope (olympus) at 100 to 400 magnifications. cladocera species were identified based on the studies by hofmann (1993b, 1999), flössner (2000), szeroczyńska and sarmajakorjonen (2007) and korosi et al. (2010). cladocera ecological preferences were defined as in flössner (2000), and błędzki and rybak (2016). the percentage composition of cladocera, the ratio of planktonic to littoral species, and the total number of identified individuals in cm³ sediment were graphically represented using c2 (juggins, 2007), while the homogeneous cladocera zones were identified using the coniss statistical method. tentative statistical analyses of cladocera results provided no significant correlation between cladocera and geochemical proxies (see drӓger et al., 2016 for further details). the exception is represented by the last ca. 100 years, which are characterized by higher trophic conditions, increasing organic carbon and calcite contents, which resulted to be related to changes in subfossil cladocera (kienel et al., 2013). the results of the morphological analysis of cladocera remains from lake tiefer see were compared with the results obtained in the 1990s for the laminated sediments of lake gościąż (with a resolution ranging from 10 to 50 years), which were published in a monographic study (ralska-jasiewiczowa et al., 1998; szeroczyńska, 1998a). results the subfossil cladocera fauna in the sediments from lake tiefer see was represented by 36 species belonging to 6 families (fig. 2). pelagic cladocera species dominated during the whole history of the lake, especially those belonging to the family bosminidae baird. littoral individuals from the family chydoridae stebbing, were represented by numerous species but showed with very low frequency (figs. 2 and 3). fig. 3 shows the comparison of long term changes in the proportion of pelagic and littoral cladoceran species in the partially varved lake tiefer see and in the completely laminated lake gościąż. as mentioned before, no significant relation was found between changes of the cladocera assemblage and geochemical sediment parameters. the exception is represented by the last ca. 100 years, which are characterized by higher trophic conditions, increasing organic carbon and calcite contents, which resulted to be related to changes in subfossil cladocera (kienel et al., 2013). while the sediment records included four distinguished units with well-varved sediments and four units with poorly or non-varved sediments (fig. 2), the coniss analysis revealed seven main cladoceran zones in relation to the species composition and the total number of cladoceran specimens, which correspond to the development phases of the lake. the first two phases corresponded to the sediment interval below 770 cm depth, while the last 6000 years are subdivided in five phases (fig. 2). in phase i (1065-970 cm) the cladocera were represented mainly by pioneer species, i.e.alonella nana baird (up to 29%), chydorus sphaericus (o.f. müller) (up to 13.5 %), alona affinis (leydig) (up to 8.3%) and acroperus harpae (baird) (up to 4.7%, fig. 2). planktonic species were dominated by forms occurring also in shallow waters, such as species from the group of daphnia longispina o.f. müller and bosmina longirostris (o.f. müller). the density of cladocera individuals gradually increased and at the end of the phase reached the number of 15,150 individuals per 1 cm3 (ind cm–3) of sediment (fig. 2). during the phase ii (970-715 cm) cladocera reached the highest density in the sediment profile, i.e. 21750 ind cm–3 at the depth of 871 cm (fig. 2). the beginning of this phase was characterized by the dominance of planktonic forms, including mainly species of the family bosminidae (figs. 2, 3 and 4). in particular the specie b. longirostris, which is a species preferring waters rich in nutrients, reached 64.5% of the total cladocera abundance while the relative abundance of the species eubosmina seligo group was over 30%. particularly noteworthy are the extreme morphs produced by species of the group bosmina (e.) longispina leydig (brooks and dodson, 1965), which were observed at the end of phase ii (fig. 5). these forms were characterised by unusual no n c om me rci al us e o nly 188 k. szeroczyńska length of mucrones on carapaces (fig. 5). phase ii was also distinguished by the increased abundance of c. sphaericus (14%), a littoral species occurring mostly in the pelagic zone, as well as of alonella excisa (fischer) (up to 4%), an acidophilous littoral species. as well as in phase ii, also in phase iii (715-584 cm, ca. 5400-4076 yr cal bp) planktonic forms were mainly represented by bosminidae with a very long mucro. moreover, this phase was characterized by the appearance of bosmina (e.) coregoni baird with very short antennae, and by high densities (up to 5.6%) of the planktonic predatory species leptodora kindti (focke). during the phase iv (584-477 cm, ca. 4076-2890 yr cal bp) the cladoceran zooplankton registered a decrease in “eutrophic” species b. longirostris and an increase in eubosmina group. extreme morphs occurred in slightly smaller numbers compared to the previous phase, which was also accompanied by the reduced count of l. kindti. noteworthy is the increased development of species of the group d. longispina (over 10%). among littoral species, monospilus dispar sars and rynchotalona falcata (sars) reached the maximum abundance, i.e. over 3% and 9.6%, respectively. cladocera reached here the lowest density in the history of lake tiefer see, which ranged from 4250 to 6550 ind cm–3. phase v (477-389 cm, ca. 2890-2055 yr cal bp) was distinguished by a significant abundance (up to 30%) of b. (e.) longispina with very long mucros (fig. 2). species of the d. longispina group also occurred in relatively large numbers, as well as the acidophilous littoral species a. excisa and graptoleberis testudinaria (fischer). phase vi (389-185 cm, ca. 2055-725 yr cal bp) was characterised by large fluctuations in the frequency of both planktonic and littoral species (fig. 3), based on which three sub-phases were distinguished (via, vib, vic; fig. 2). frequency of cladocera specimens increased from 5150 ind cm–3 in via to 16,350 ind cm–3 in vib and only a very few extreme morphs were found in these subphases. subphase vic was characterised by the presence of eubosmina species with very short antennae, and by bosmina (e.) reflexa seligo, which was identified for the first time in this layers. pleuroxus leavis sars and c. sphaericus reached in via and vic a considerable abundance over 6% in the littoral zone, while the total number fig. 3. comparison of the percentage content of planktonic and littoral cladocera species from the sediments of lake tiefer see (germany) and lake gościąż (poland). 1, well-varved section (regular laminae); 2, non-varved section; 3, section irregularly laminated. no n c om me rci al us e o nly long term subfossil cladocera from lake tiefer see 189 of all individuals amounted to more than 16,000 ind cm– 3 in sub-phase vib (fig. 2). phase vii (185-0 cm, ca. 725 yr cal bp to modern time) was characterised by large fluctuations in the species frequency, on the basis of which three sub-phases were distinguished (fig. 2). phase vii registered a re-occurrence of extreme morphs of b. (e.) longispina type with a very long mucro, while the “eutrophic” species b. longirostris increased significantly, reaching over 80% of the relative abundance in sub-phase viib. sub-phase viia highlighted an overall decline in the total abundance of cladocera (fig. 2), which was accompanied by the complete disappearance of certain species (e.g. alonella sars, pleuroxus baird, alona baird) and a significant contribution of c. sphaericus (up to 0.9%). it was observed that planktonic species, especially taxa preferring nutrient-rich waters, developed again at the end of the phase (viic), with b. longirostris (ca. 65%) as a dominant species. reconstructed vegetation showed higher openness during the periods ca. 3900-3100, 2700-2200 and 750 yr cal bp (fig. 4; see drӓger et al., 2016 for further details). discussion phases of the lake development the analysis of species composition and changes in the frequency of subfossil cladocera represented in the lacustrine sediments of lake tiefer see allowed the reconstruction of the lake evolution. on the basis of the characteristics of cladocera assemblages in lake tiefer see seven main phases of lake development during the holocene were distinguished. phase i corresponded to the initial period of the lake evolution, which is likely attributable to the end of the late glacial period and the early holocene (ralskajasiewiczowa, 1998; brauer et al., 1999). autochthonous lacustrine non–varved sediments were deposited during this stage. the presence of cladocera pioneer species occurring in association with aquatic vegetation (a. harpae, a. affinis) indicates lower water level or/and pronounced re-deposition of sediment from the littoral zone of the lake. the occurrence of planktonic species (eubosmina, fig. 4. comparison between stages of trophic increase (dotted vertical grey arrows) and reconstructed vegetation openness (as an index of settlements phases). sediment column with varved (units: i, iii, v, vii) and non-varved (units: ii, iv, vi) sections after dräger et al. (2016). 1, well-varved section (regular laminae); 2, non-varved section; 3, section irregularly laminated. no n c om me rci al us e o nly 190 k. szeroczyńska and d. longispina group) and the high total number of cladocera at the end of this phase indicated more stable conditions, and that water temperature and edaphic conditions progressively became very favourable for the zooplankton growth (frey, 1986; jeppesen et al., 2000; korosi and smol, 2012). such a picture is often found in lacustrine sediments deposited during the transition between the late glacial and the early holocene period (szeroczyńska, 1998a, 2006; zawisza and szeroczyńska, 2007; kulesza et al., 2011; zawiska et al., 2015). the species composition of subfossil cladocera during phase ii indicated optimal lake environmental conditions. during this time well-varved sediments were deposited (with two gaps of sediments in the composite profile). the increased abundance of planktonic [in particular b.(e.) longispina and b.(e.) coregoni], littoral acidophilous species (i.e.chydorus piger sars, a. excisa), together with the decrease of species living in association with aquatic vegetation (e.g., a. harpae, a. affinis) indicate alpha/or beta-mesotrophic status and higher water level (alhonen, 1970; sarmaja-korjonen, 2001; korosi and smol, 2012). probably the sediments of phase ii were mainly deposited during the holocene climatic optimum. the growth of cladocera fauna was obviously affected by warm and fig. 5. remains of eubosmina extreme morphs. 1-4: shells with very long mucros; 5-6: head shield with very short antennas head pores eubosmina coregoni type. scale bars: 100 µm. no n c om me rci al us e o nly long term subfossil cladocera from lake tiefer see 191 humid climate prevailing in the atlantic period. noteworthy is the high abundance of b. longirostris, a species preferring waters with a higher trophic status, as it confirms the high lake productivity (szeroczyńska, 1998b; jeppesen et al., 2000; manca et al., 2007; de sellas et al., 2008). phase iii probably coincided with the first half of the subboreal period. the deposition of the well-varved sediments continued in the lake during this stage. after the holocene climate optimum (ca. 8000 to 3500 cal yr bp), during which warm temperature and edaphic conditions prevailed, both planktonic and littoral cladocera decreased. the observed morphological modifications, i.e. changes in the shape and size of bosminidae individuals, indicated that either water level oscillations and/or high predation pressure from both invertebrates and fish, as previously outlined by several studies conducted at the lake (e.g. kerfoot, 1981; korosi et al., 2008, 2010, 2013; swetmann and finney, 2003; sakamoto et al., 2007; sakamoto and hanozato, 2008; błędzki and szeroczyńska, 2015). the climate conditions probably changed during phase iv and became cool (windy) and humid. such a picture has been often observed during the subboreal to subatlantic climate transition (milecka and szeroczyńska 2005; zawisza and szeroczyńska 2007). probably water mixing and water-level fluctuations increased at that time (drӓger et al., 2016), thus enhancing water turbulence and turbidity and preventing both a stable deposition of sediments and the development of zooplankton. in fact non-varved sediments were deposited during this time. higher water turbidity (low transparency) and circulation are generally considered as not conducive to biological production including the growth of cladocera fauna (cottenie and de meester, 2003; korosi et al., 2013). such environmental conditions are manifested by changes in the frequency of daphnia individuals as well as of littoral species. during this phase, lake waters were also probably relatively poor in nutrients, which most likely contributed to the significant drop in the abundance of all cladocera species. r. falcata, which is a littoral species frequently present in lobelia-type lakes, occurred in large numbers at that time, thus indicating oligotrophic conditions (milecka and szeroczyńska, 2005; błędzki and rybak, 2016). phase v coincided with the initial phase of the humid and relatively colder subatlantic period (after ca. 2600 yr cal bp). during this phase poorly varved sediment was deposited and the total abundance of cladocera was only slightly higher compared to phase iv. this period was characterised by the high contribution of extreme forms. morphological modifications of bosminidae not only protect them against predators, but may also reduce the effectiveness of swimming and consequently increase the resistance of animals against unfavourable hydrodynamics of water (gliwicz and pijanowska, 1989; gliwicz et al., 2000; korosi et al., 2013). perhaps stronger wave motion and frequent water mixing induced an increased production of these extreme morphs. in the second part of the phase, since ca. 2400 yr cal bp, environmental conditions in the lake considerably improved. bosminidae produced only “normal morphs”, and the abundance of plankton increased. the climate became warmer and the water temperature became conducive to the development of the fauna (gliwicz, 1990; günter and lieder, 1993; jeppesen et al., 2000; korosi et al., 2013). after a fairly long period of relatively adverse lake conditions, edaphic and temperature conditions improved during the phases vi. well varved sediment (unit vi) was deposited during the warmer part of subatlantic period, and the total abundance of zooplankton increased. both pelagic and littoral cladocera species experienced favourable conditions for their development. during this period, bosminidae produced only rare extreme morphs, which may indicate a reduced pressure of predators and/or more stable hydrological conditions (jeppesen et al., 2000; nevalainen and luoto, 2016). the water temperature was optimal for the development of most cladocera species (e.g., p. laevis, camptocercus rectirostris schoedler), indicating favourable climate conditions (błędzki and rybak, 2016). the record of species indicating increased nutrient availabilities (i.e. alonella exigua (lilljeborg), disparalona rostrata (koch), pleuroxus uncinatus baird), probably reflects short-term fluctuations in the trophic conditions of the water, which might reflect the first human impact on the lake tiefer see ecosystem. in fact vegetation openness reconstructed from pollen data has been related to human activity within the lake catchment (dräger et al., 2016) during phases vii, which spans over the last millennium, non-varved (in subphases viia-b) and well varved sediments (unit viic) were deposited. the species composition and low cladocera density indicate that during this part of the subatlantic period the conditions in the lake were very unstable, thus confirming the types of sediment depositions. the production extreme morphs of bosmina was supposed to be related to water mixing and water level fluctuations (hellsten and stenson, 1995; lord et al., 2006). the reduced frequency and disappearance of certain species was observed in the sediment deposited ca. 640-440 yr cal bp, and ca. 300 years ago. these changes might be interpreted as the result of both climate cooling and reduction in the lake nutrient level during the little ice age (drӓger et al., 2016). the conditions in the lake changed again during the deposition of the youngest sediments (varved sediment unit in phase vii). the occurrence of taxa indicating higher nutrient availability may indicate an inflow of nutrients into the lake and an increased the zooplankton production both in the pelagic and littoral zone. the characteristics of both the open water and littoral no n c om me rci al us e o nly 192 k. szeroczyńska cladoceran fauna along the sediment core from lake tiefer see indicated the lake basin as deep during its whole history, and with a well-developed pelagic zone, as indicated by the dominating planktonic species. the most suitable conditions for the development of water fauna occurred during phases ii and vi, which well correspond to the deposition of well-varved sediments. on the contrary, lower frequency of cladoceran individuals was observed during stages of nonand poorly varved sediment deposition (especially cladocera phase iv and sub-phase viia). a different situation was noted in the non-varved sub-phase viib, where the total number of cladoceran increased, due to the high density of b. longirostris, i.e. a species preferring high nutrient availability. this indicates a recent increase in the lake trophic status in relation to anthropogenic impact. the lower cladocera specimen number in poorlyand non-varved intervals in contrast to well-varved sediment phases implies that increased lake circulation might have negatively affected cladoceran populations in the lake, as observed by jeppesen et al., (2000) and cottenie and de meester (2003). the planktonic b. longirostris and the littoral c. sphaericus occurred with high abundance in lake tiefer see. both species are often dominant in eutrophic freshwater bodies (alhonen, 1970; hofmann, 1996; szeroczyńska, 1998a, 1998b; korhola and rautio, 2001; sarmaja-korjonen, 2001; schmidt et al., 2001). however, taking into account the high occurrence of other pelagic species, which prefer lakes with lower trophy, it can be presumed that lake tiefer see was never extremely eutrophic, and the lake oscillated during its evolution between alphaand beta-mesotrophic status. eight stages of increasing trophy, indicated by species associated with high lake productivity, were identified, in particular between 5750-5500, 4500-4100 and 750-50 yr cal bp. increases in the nutrient availability were likely more related to climatic conditions during the early holocene, and to combined climate and anthropogenic factors during the late holocene. increasing abundance of b. longirostris indicated higher productivity and an increased contribution of this species to the cladoceran fauna of lake tiefer see was also observed in earlier periods, especially during the first stage of the lake development, which was accompanied by a considerable growth of littoral species. sediments of many lakes (both deeper and shallow, lowland and mountain lakes) show such a picture in the early holocene (szeroczyńska, 1998a, 2006), when climate warming, and hence the rise in water temperature, provided good conditions for the development of flora and fauna. the second major increase in the abundance of b. longirostris in lake tiefer see was observed at a depth of ca. 900 cm, which was accompanied by the maximum increase in the number of species individuals. such a picture is often observed during the holocene climate optimum when the plankton, reached the maximum growth especially in deeper lakes (milecka and szeroczyńska, 2005). the increased trophic status of the lake observed during the period from approximately 6000 yr cal bp to the modern times was mainly the effect of anthropogenic activity, rather than climate. however, not all periods of anthropogenic activity are reflected by changes in productivity of lake tiefer see, such as phase iv and v, non varved sediments units ii and iv. a probable cause might be the increased circulation in the lake during these intervals (restricted transmission of light) which significantly influences cladocera growth. comparable trophic changes in response to human impact, i.e. during the periods 5850-5450, 4400-3800, 19501700 yr cal bp, have been reconstructed on the basis of subfossil cladocera analyses in lake gościąż (ralskajasiewiczowa et al., 1998; szeroczyńska, 1998a). similarly as in lake tiefer see, the highest increase in the trophic status occurred in the period from 750 yr cal bp until the modern times. generally lake gościąż was classified as meromictic, mesotrophic, while tiefer see is dimictic, and maintained an oligoor mesotrophic character during its development. at present lake gościąż, unlike lake tiefer see, is classified as eutrophic. despite the fact that lake gościąż is even more isolated from the direct human impact, the process of eutrophication is much faster compared to lake tiefer see. this is primarily due to its much smaller depth and constant stratification (meromixis). cladocera morphology interestingly, different morphological types of b. longirostris, b.(e) coregoni and b.(e) longispina were observed along the sediment profile. forms of b. (e) coregoni showed pronounced variability of mucro length on shell, antennae length, together with the head shield. all identified forms of the species and its subspecies had head pores on the head shield, which are characteristic of b. (e) coregoni types. other lake sediment studies conducted in europe have outlined the presence of species of eubosmina with different morphological characteristics, such as b. (e) coregoni with a short or no mucro and long antennae, b. (e) longispina and b.(e) reflexa with a mucro of varying size (hofmann, 1996; gasiorowski and szeroczyńska, 2004; faustova et al., 2011; błędzki and szeroczyńska, 2015). as most of the lakes, where these forms have been identified, are mesotrophic, it may be assumed that trophic conditions and water temperature are the decisive factors for the existence of different eubosmina subspecies. hofmann (1993b; 1996) determined high diversity of eubosmina by measuring the mucro length and determined that the variability of eubosmina species was evolutionary. in his papers on lakes grosser segebergersee, grosser plönersee and bodensee (lake constance), hofmann (1984) described cladoceran species that are characteristic of a given climate period. no n c om me rci al us e o nly long term subfossil cladocera from lake tiefer see 193 different morphological characteristics of the species b. (e) longispina were present in late glacial times, of bosmina (e) coregoni f. kessleri uljanin in the boreal and atlantic period, and of b. (e) coregoni in the youngest subboreal and subatlantic period. however, in the non varved polish lakes ostrowite and charzykowskie (milecka and szeroczyńska, 2005; błędzki and szeroczyńska, 2015) and in lake tiefer see (germany), an alternating (non-evolutionary) occurrence of different eubosmina forms was found at different sediment depths. despite the similarities between the two varved lakes tiefer see and gościąż, the latter showed a pronounced scarcity in the abundance of bosminidae and, in particular, the lack of extreme morphs. this raises an important question of why there were so many morphs in lake tiefer see and what was the main cause of their increased production. furthermore, it is still not known whether the development of different morphs is connected with thermal or other physical conditions, or to water chemical changes. the results from the analysed lakes in poland and germany allow the assumption that the change in lake productivity was the driver for morphological variability of cladocera. in poland, subfossil cladocera were analysed mainly in lakes with higher trophic status, from mesotrophic to hypertrophic, while studies on sediments from deep oligotrophic mountain lakes revealed a practically complete absence of species belonging to the bosminidae family. oligotrophic lakes in the tatra mountains (szeroczyńska, 2006; sienkiewicz and gąsiorowski, 2016), probably had such a low availability of nutrients during the whole holocene, that edaphic conditions were almost never suitable for the bosminidae existence. b. longirostris was found only in the sediment deposited during the modern time. it seems that the occurrence of bosmina, especially b.(e) coregoni and its morphological variants, is mainly related to edaphic conditions (cottenie and de meester, 2003), and it can be assumed that the mesotrophic state of waters is favourable for the development of some variations of eubosmina (hofmann, 1996; korosi et al., 2010). probably the mesotrophic status of lake tiefer see provided good conditions for abundant occurrence of planktonic species, but the question still remain on what factors can induce the development of these varieties, including extreme morphs? why eubosmina morphs characterised by a very small head shield with very short antennae occurred only in varved sediment intervals, while morphs with very long mucros dominated in non-varved sediment sections? such forms were also found in varved sediments deposited in the period 55004000 yr cal bp. for years the occurrence of different morphs of bosminidae represented a problematic issue for the researchers (brooks and dodson, 1965; kerfoot, 2006; sakamoto and hanazato, 2008; błędzki et al., 2013; korosi et al., 2013; błędzki and szeroczyńska, 2015). most researchers suggest that the main reason for producing morphological variations is the adaptation of the species to extreme conditions. the variations in size and the production of different morphs protect the individuals from predation. in the case of bosminidae, it provides protection from predation by other invertebrates. in fact it has been shown that bosminidae often reduces their body size to get protected from predation by copepoda (sakamoto and hanazato, 2008). another adaption consists in a significant increase of the body size (e.g., extending antenna or mucro) to reduce the risk of predation from other cladocera, such as l. kindti. these morphological changes could be observed in lake tiefer see, where the increased frequency of b. (e.) longispina with a very long mucro correlates with high relative abundance of l. kindti. through the mechanism of direct identification of predators, also daphniidae straus can significantly reduce or increase their body size by producing high helmets and long spines to be protected from fish predation. however, not only their ability to detect the threat from predators, but also environmental factors may cause morphological changes. the increase in water temperature as well as seasonal and chemical changes can also support the formation of different morphs (gliwicz, 1990; moore and folt, 1993; sakamoto and hanazato, 2008; korosi et al., 2013). the modified morphology can reduce the swimming efficiency and increase the resistance of animals to water turbulence. this seems to be the probable explanation of extreme morphs found in varved lake sediments and other morphs found in non-varved sediments in lake tiefer see. it can be supposed that in periods of non-varved sedimentation, the lake water was less transparent and more turbulent, what induced bosminidae to produce morphs provided with more physical stability and slower movement (gliwicz and pijanowska, 1989). this caused no disadvantage in the case of homogenous distribution of the food due water turbulence. thus, morphological variations support the survival of individuals in unstable and less favourable biotic and abiotic conditions, which include pressure by predators (invertebrates and fish), temperature, strong water movements and rapid chemical variability. conclusions the species composition of cladocera determined in varved and non-varved sediments showed that lake tiefer see was a deep water body throughout the holocene. the lake was often exposed to increased water mixing, especially after 4000 yr cal bp, and this was reflected by the deposition of varved and non-varved sediments and by the production of extreme morphs by the cladoceran species of the family bosminidae. the comno n c om me rci al us e o nly 194 k. szeroczyńska parison with other studies on subfossil cladocera in long sediment cores suggests that the production of morphologically variable forms was the mechanism that enabled the zooplanktonic organisms to adapt to changing biotic (pressure of predators) and abiotic conditions (temperature, turbid water). the best lacustrine conditions for the development of fauna occurred during the sedimentation of well varved sediments, which is reflected by the increase of abundance of all species, which are indicators of higher nutrient availability. increases in the zooplankton frequency and, at the same time, in the lake trophic status coincided with the periods of climate warming in the early and middle holocene and with the impact of human colonisation during successive phases from the late neolithic to present day. the comparison of subfossil cladocera evolution during the entire holocene in two lakes with laminated sediments, i.e. lake tiefer see (germany) and lake gościąż (poland), outlined significant differences which were responsible for the different cladocera-inferred reconstruction of the long term trophic evolution of the two lakes. this is noteworthy when considering that the two lakes are located in a similar geographical zone, that both are influenced by oceanic climate regime, and that they are scarcely affected by direct human impact. the meromixis was likely a key factor for the evolution of lake gościąż, whereas the frequent water mixing stages were determinant for lake tiefer see. the results provided by the investigation of annually laminated sediments excellent reference material for reconstructing of the past cladoceran dynamics and lacustrine environments, as well as for predicting future ecological trends especially in connection of human impact. acknowledgments this study is a contribution to the virtual institute of integrated climate and landscape evolution analysis (iclea) of the helmholtz association (grant number vh-vi-415) and used infrastructure of the terrestrial environmental observatory (tereno). furthermore, this research was possible with a support of the institute of geological sciences, polish academy of sciences. i wish to thank achim brauer and nadine drӓger (gfz german research centre for geosciences, potsdam, germany) for providing sediment samples and additional data for this review. i am also thankful to all technicians, who helped in the laboratory and subfossil cladocera analysis, to edyta zawisza and nadine drӓger for valuable comments on the original manuscript, and to two anonymous reviewers and the editors for very constructive remarks to improve this manuscript. references alhonen p, 1970. on the significance of the planktonic/littoral ratio in the cladoceran stratigraphy of lake sediments. comm. biol. 35:1-9. bennion h, battarbee rw, sayer cd, simpson gl, davidson ta, 2011. defining reference conditions and restoration targets for lake ecosystems using palaeolimnology: a synthesis. j. paleolimnol. 45:533-544. błędzki la, rybak ji, 2016. freshwater crustacean zooplankton of europe. springer, switzerland: 918 pp. błędzki la, szeroczyńska k, puusepp e, 2013. the late holocene appearance of european bosmina (eubosmina) thersites (crustacea, cladocera) in lakes surrounding the baltic sea. hydrobiologia 715:77-86. błędzki la, szeroczyńska k, 2015. palaeolimnological evidence of bosmina morphotypes appearance in the late holocene. holocene 25:557-561. boucherle mm, züllig h, 1983. cladoceran remains as evidence of change in trophic state in three swiss lakes. hydrobiologia 103:141-146. brauer a, 2004. annually laminated lake sediments and their palaeoclimatic relevance, p. 109-127. in: h. fischer, t. kumke and g. lohmann g (eds.), the climate in historical times. springer, berlin. brauer a, enders c, negendank jfw, 1999. lateglacial calendar year chronology based on annually laminated sediments from lake meerfelder maar, germany. quat. int. 61:17-25. brauer a, casanova j, 2001. chronology and depositional processes of the laminated record from lac d’annecy, french alps. j. paleolimnol. 25:163-177. brooks jl, dodson si, 1965. predation, body size, and composition of plankton. science 150:28-35. chen g, dalton c, taylor d, 2010. cladocera as indicators of trophic state in irish lakes. j. paleolimnol. 44:465-481. cottenie k, de meester l, 2003. connectivity and cladoceran species richness in a metacommunity of shallow lakes. freshwater biol. 48:823-832. de sellas am, paterson am, sweetman jn, smol jp, 2008. cladocera assemblages from the surface sediments of southcentral ontario (canada) lakes and their relationships to measured environmental variables. hydrobiologia 600:105-119. dräger n, theuerkauf m, szeroczyńska k, wulf s, tjallingii r, plessen b, kienel u, brauer a, 2016. varve microfacies and varve preservation record of climate change and human impact for the last 6000 years at lake tiefer see (ne germany). holocene doi: 101177/0959683616660173. faustova m, sacherova v, svensson je, taylor dj, 2011. radiation of european eubosmina (cladocera) from bosmina (e.) longispina – concordance of multipopulation molecular data with paleolimnology. limnol. oceanogr. 56:440-450. flössner d, 2000. [die haplopoda und cladocera mitteleuropas].[book in german]. backhuys publishers, leiden: 440 pp. frey dg, 1986. cladocera analysis, p. 667-692. in: b.e. berglund (ed.), handbook of holocene palaeoecology and palaeohydrology. j. wiley & sons, chichester. gąsiorowski m, szeroczyńska k, 2004. abrupt changes in bosmina (cladocera, crustacea) assemblages during the history of the ostrowite lake (northern poland). hydrobiologia no n c om me rci al us e o nly long term subfossil cladocera from lake tiefer see 195 526:137-144. gliwicz zm, 1990. food thresholds and body size in cladocerans. nature 343:638-640. gliwicz zm, pijanowska j, 1989. the role of predation in zooplankton succession, p. 253-296. in: u. sommer (ed.), plankton ecology: succession in plankton communities. springer, heidelberg. gliwicz zm, rutkowska ae, wojciechowska j, 2000. daphnia populations in three interconnected lakes with roach as the principal planktivore. j. plankton res. 22:1539-1557. günther j, lieder u, 1993. postglacial succession in the subgenus eubosmina (crustacea: cladocera) in the region of the unterhavel river (near berlin, germany) type changes or species immigration? int. rev. ges. hydrobiol. 78:1-19. hall ri, smol jp, 1996. paleolimnological assessment of longterm water quality changes in south-central ontario lakes affected by cottage development and acidification. can. j. fish. aquat. sci. 53:1-17. hann bj, leavitt pr, chang pss, 1994. cladocera community response to experimental eutrophication in lake 227 as recorded in laminated sediments. can. j. fish. aquat. sci. 51:2312-2321. hellsten ae, stenson jae, 1995. cyclomorphosis in a population of bosmina coregoni. hydrobiologia 312:1-9. hofmann w, 1984. postglacial morphological variation in bosmina longispina leydig (crustacea, cladocera) from grosser plöner see (north germany) and its taxonomic implications. z. zool. syst. evol. -forsch. 22:294-301. hofmann w, 1993a. late-glacial / holocene changes of the climatic and trophic conditions in three eifel maar lakes, as indicated by faunal remains. i. cladocera. lect. notes in earth sci. 49:393-420. hofmann w, 1993b. morphological variation in the planktonic cladoceran bosmina (eubosmina) in the selenter see. faunist.-ökol. mitt. 6:479-485. hofmann w, 1996. empirical relationships between cladoceran fauna and trophic state in thirteen northern german lakes: analysis of surficial sediments. hydrobiologia 318:195-201. hofmann w, 1999. holocene succession and morphological variation of the bosmina (eubosmina) taxa of the plußsee (northern germany). arch. hydrobiol. spec. issues advanc. limnol. 54:359-372. hofmann w, 2001. late-glacial/holocene succession of the chironomid and cladoceran fauna of the soppensee (central switzerland). j. paleolimnol. 25:411-420. jeppesen e, jensen jp, sǿndergaard m, lauridsen t, landkildehus f, 2000. trophic structure, species richness and biodiversity in danish lakes; changes along phosphorus gradient. freshwater biol. 45:201-218. juggins s, 2007. c2 ver. 1.5 user guide. software for ecological and palaeoecological data analysis and visualisation. newcastle university, newcastle upon tyne, uk 73. kamenik c, szeroczyńska k, schmidt r, 2007. relationships among recent alpine cladocera remains and their enviroment: implications for climate-change studies. hydrobiologia 595:33-46. kemp aes, 2016. laminated sediments as paleo-indicators. in: a.e.s. kemp (ed.), palaeoclimatology and palaeoceanography from laminated sediments. geological society special publication no. 116:70-120. kerfoot wc, 1981. long-term replacement cycles in cladoceran communities: a history of predation. ecology 62:216-233. kerfoot wc, 2006. baltic eubosmina morphological radiation: sensitivity to invertebrate predators (induction) and observations on genetic differences. arch. hydrobiol. 167:147-168. kienel u, dulski p, ott f, lorenz s, brauer a, 2013. recently induced anoxia leading to the preservation of seasonal laminae in two ne-german lakes. j. paleolimnol. 50:535-544. kinder m, tylman w, enters d, piotrowska n, poręba g, zolitschka b, 2013. construction and validation of calendaryear time scale for annually laminated sediments an example from lake szurpity (ne poland). gff 135:248-257. korhola a, rautio m, 2001. cladocera and other brachiopod crustaceans, p. 5-41. in: j.p. smol, h.j.b. birks and w.m. last (eds.), tracking environmental change using lake sediments. 4. zoological indicators. kluwer, dordrecht. korosi jb, paterson am, desellas am, smol jp, 2008. linking mean body size of pelagic cladocera to environmental variables in precambrian shield lakes: a paleolimnological approach. j. limnol. 67:22-34. korosi jb, paterson am, desellas am, smol jp, 2010. a comparison of pre-industrial and present-day changes in bosmina and daphnia size structure from soft-water ontario lakes. can. j. fish. aquat. sci. 67:754-762. korosi jb, smol jp, 2012. contrasts between dystrophic and clearwater lakes in the long-term effects of acidification on cladoceran assemblages. freshwater biol. 57:2449-2464. korosi jb, kurek j, smol jp, 2013. a review on utilizing bosmina size structure archived in lake sediments to inter hidtoric shift in predation reims. j. plankton res. 35:444-460. kulesza p, suchora m, pidek ia, alexandrowicz wp, 2011. chronology and directions of late glacial paleoenvironmental changes: a multi-proxy study on sediments of lake slone (se poland). quater. int. 238:89-106. last wm, smol jp, 2001. tracking environmental change using lake sediments. 2. physical and geochemical methods. kluwer, dordrecht: 504 pp. lord h, lagergren r, svensson je, lundqvist n, 2006. sexual dimorphism in bosmina: the role of morphology, drag, and swimming. ecology 87:788-795. manca m, comoli p, 1996. reconstructing population size structure in cladocera by measuring their body remains. mem. ist. ital. idrobiol. 54:61-67. manca m, torretta b, comoli p, amsinck sl, jeppesen e, 2007. major changes in trophic dynamics in large, deep, sub-alpine lake maggiore from 1940s to 2002: a high resolution comparative palaeo-neolimnological study. freshwater biol. 52:2256-2269. milan m, 2016. long-term development of subalpine lakes: effects of nutrients, climate and hydrological variability as assessed by biological and geochemical sediment proxies. ph.d. thesis, umeå university, sweden. milecka k, szeroczyńska k, 2005. changes in macrophytic flora and planctonic organisms in lake ostrowite, poland, as a response to climatic and trophic fluctuactions. holocene 15:74-84. mingram j, negedank jfw, brauer a, berger d, hendrich a, köhler m, unsinger h, 2007. long cores from small lakes recovering up to 100 m-long lake sediment sequences with a high-precision rod-operated piston corer (unsiger-corer). no n c om me rci al us e o nly 196 k. szeroczyńska j. paleolimnol. 37:517-528. mirosław-grabowska j, zawisza e, 2014. late glacial-early holocene environmental changes in charzykowskie lake (northern poland) based on oxygen and carbon isotopes and cladocera data. quarter. int. 328/329:156-166. moore m, folt c, 1993. zooplankton body size and community structure: effects of thermal and toxicant stress. trends ecol. evol. 8:178-183. nevalainen l, luoto tp, 2016. relationship between cladoceran (crustacea) functional diversity and lake trophic gradients. funct. ecol. doi:10.1111/1365-2435.12737. niska m, mirosław-grabowska j, 2015. eemian environmental changes recorded in lake deposits from rzecino (nw poland): cladocera, isotopic and selected geochemical data. j. paleolimnol. 53:89-105. nykänen m, malinen k, liukkonen m, kairesalo t, 2010. cladoceran community responses to biomanipulation and re-oligotrophication in lake vesijärvi, finland, as inferred from remains in annually laminated sediment. freshwater biol. 55:1164-1181. ralska-jasiewiczowa m, goslar t, madeyska t, starkel l (eds.), 1998. lake gościąż, central poland a monographic study, part i w. szafer institute of botany, krakow: 340 pp. sakamoto m, kwang-hyeon c, hanazato t, 2007. plastic phenotypes of antennule shape in bosmina longirostris controlled by physical stimuli from predators. limnol. oceanogr. 52: 2072-2078. sakamoto m, hanazato t, 2008. antennule shape and body size of bosmina: key factors determining its vulnerability to predacious copepoda. j. limnol. 9:27-34. sarmaja-korjonen k, 2001. correlation of fluctuations in cladoceran planktonic: littoral ratio between three cores from a small lake in southern finland: holocene water-level changes. holocene 11:53-63. schmidt r, muller j, drescher-schneider r, szeroczyńska k, barri a, krisai r, 2001. changes in holocene lake level and production in a large northern adriatic karstic lake (lake vrana, cres, croatia). terra nostra 1:53-56. sienkiewicz e, gąsiorowski m, 2016. the effect of fish stocking on mountain lake plankton communities identified using palaeobiological analyses of bottom sediment cores, j. paleolimnol. 55:129-150. sweetman jn, finney bp, 2003. differential responses of zooplankton populations (bosmina longirostris) to fish predation and nutrient loading in an introduced and natural sockeye salmon nursery lake on kodiak island, alaska, usa. j. paleolimnol. 30:183-193. szeroczyńska k, 1998a. the holocene cladoceran succession in the laminated sediments of lake gościąż, p. 219-225. in: m. ralska-jasiewiczowa, t. goslar, t. madeyska and l. starkel (eds), lake gościąż, central poland a monographic study, part i. w. szafer institute of botany, krakow. szeroczyńska k, 1998b. anthropogenic transformation of nine lakes in central poland from mesolithic to modern times in the light of cladocera analysis. stud. geol. pol. 112:123-165. szeroczyńska k, 2006. the significance of subfossil cladocera in stratigraphy of late glacial and holocene. stud. quarter. 23:37-45. szeroczyńska k, sarmaja-korjonen k, 2007. atlas of subfossil cladocera from central and northern europe. friends of the lower vistula society, świecie: 84 pp. sturm m, 1979. origin and composition of plastic varves, p. 281-285. in: c. schlüchter (ed.), moraines and varves. a.a. balkema, rotterdam. sturm m, lotter af, 1995. lake sediments as environmental archives. ewag news 38:6-9. tylman w, szpakowska k, ohlendorf c, woszczyk m, zolitschka b, 2012. conditions for deposition of annually laminated sediments in small meromictic lakes: a case study of lake suminko (northern poland). j. paleolimnol. 47:55-70. tylman w, zolitschka b, enters d, olehndorf c, 2013. laminated lake sediments in northeast poland: distribution, preconditions for formation and potential for paleoenvironmental investigation. j. paleolimnol. 50:487-503. van damme k, kotov aa, 2016. the fossil record of the cladocera (crustacea: branchiopoda): evidence and hypotheses. earth sc. rev. 163:162-189. wulf s, dräger n, ott f, serb j, appelt o, guðmundsdóttir e, van den bogaard c, słowiński m, błaszkiewicz m, brauer a, 2016. holocene tephrostratigraphy of varved sediment records from lakes tiefer see (ne germany) and czechowskie (n poland). quater. sc. rev. 132:1-14. zawiska i, słowiński m, correa-metrio a, obremska m, luoto t, nevalainen l, woszczyk m, milecka k, 2015. the response of a shallow lake and its catchment to late glacial climate changes a case study from eastern poland. catena 126:1-10. zawisza e, szeroczyńska k, 2007. the development history of wigry lake as shown by subfossil cladocera. geochronometria 27:67-74. zolitschka b, 2007. varved lake sediments, p. 3105-3114. in: s.a. elias (ed.), encyclopedia of quaternary sciences. elsevier. zolitschka b, francus p. ojala aek, schimmelmann a, 2015. varves in lake sediments a review. quaternary sci. rev. 117:1-41. no n c om me rci al us e o nly layout 1 introduction the increasing trophic status or eutrophication of water bodies is a modern worldwide problem. nowadays, rapid increases of lake trophic levels can be easily observed in a large number of lake ecosystems (schmidt et al., 2000; richerson et al., 2008). the natural dynamics of many lake ecosystems, especially those located in urban and agricultural areas, have been highly modified by human activity. the effects of human related nutrient supply have been widely investigated and discussed (joye, 2006; smith and schindler, 2009; gałka et al., 2014). it is common opinion that major human impacts on lake trophic processes began in the middle age or in the era of the industrial revolution (starting from the 19th century). however, paleolimnological studies have demonstrated that ancient cultures with a lesser technological development, such as the neolithic or even older mesolithic human groups, were also able to substantially impact lake ecosystems (goransson, 1986; latałowa, 1994, 2003). in addition, changes in lake trophic level have also resulted from natural processes (mostly related to climate changes), and from their interactions with anthropogenic influence (zawisza and szeroczyńska, 2007; nevalainen et al., 2013; grabowska et al., 2015). as a result, the assessment of the relative influence of human and natural impacts on lakes in past times and at present is critical for establishing viable management plans for these ecosystems. in this sense, paleolimnological studies have substantially contributed to understanding the long term dynamics of eutrophication processes, e.g. in discriminating the origin of nutrients (human or natural). of special interest is the assessment of the time span required by impacts and processes to act upon water bodies and produce advances in oceanography and limnology, 2016; 7(2): 163-176 article doi: 10.4081/aiol.2016.6234 this work is licensed under a creative commons attribution-noncommercial 4.0 international license (cc by-nc 4.0). subfossil cladocera and pollen as indicators of natural and anthropogenic trophic changes of lake jelonek (tuchola forest, n poland) during the holocene edyta zawisza,1* anna filbrandt-czaja,2 alexander correa-metrio3 1institute of geological sciences, polish academy of sciences, research centre in warsaw, ing pan, twarda 51/55, pl-00818 warsaw, poland; 2chair of geobotany and landscape planning, faculty of biology and environmental protection, nicolaus copernicus university in toruń, lwowska 1, toruń pl-87100, poland; 3instituto de geología, universidad nacional autónoma de méxico, ciudad universitaria, 04510 d.f., mexico *corresponding author: ezawisza@twarda.pan.pl abstract lake jelonek is a small lake located in central northern poland, in the tuchola forest. the sediments of the lake represent a natural archive that offers insights into the natural history of the region from the late glacial to present. in winter 2002, a 1330 cm long sediment core was recovered from the deepest part of lake. using a multiproxy approach (cladocerans, pollen and basic geochemistry), we reconstructed trophic status changes through the last ~15,000 years. special attention was devoted to the evaluation of nutrient contributions to the lake from natural and anthropogenic sources. the cladocera analyses yielded a total of 29 species belonging to five families (bosminidae, daphniidae, leptodoridae, chydoridae, sididae), with planktonic species representing more than 60% of cladocera relative abundance throughout the core. the pollen results suggested four periods of increased human activity, so-called settlement phases. the first traces of human activity in the basin of lake jelonek appeared in the atlantic period and were related with mesolithic and neolithic settlements. the second (bronze age) and the third (iron age) settlement phases are well marked by the paleolimnological proxies studied. this time period clearly manifested on the lake waters as an increasing trophy level probably caused by human-associated discharges of nutrients to the lake. after the third settlement phase cladoceran data indicated a significant decrease in the lake trophic level and the pollen data showed a recovery of forest cover. the fourth period of human economic activity during the early middle age was characterized by deforestation associated with land reclamation for grazing and cultivation of cereals, and the subsequent nutrient enrichment of lake waters. according to our results, the biological development of lake jelonek was determined by climate changes from late glacial up to the atlantic period. contrastingly, the most important driver for the lake development though the middle and late holocene has been human activity within the lake catchment. the recovery of the lake trophic level during the last century, which is primarily related to extensive afforestation of the lake catchment, highlighted the importance of land use in the conservation of healthy water bodies key words: trophic changes; subfossil cladocera; pollen analysis; tuchola forest; paleolimnological multi-proxy analysis; holocene. received: august 2016. accepted: november 2016. no n c om me rci al us e o nly 164 e. zawisza et al. significant changes in their trophic status. it is well-known that lake-catchment systems respond physically, chemically and biologically to natural climatic changes as well as to environmental changes caused by human activities (o’sullivan, 1983; ralska-jasiewiczowa et al., 2003). lake ecosystems most rapidly and clearly respond to natural or anthropogenic nutrient supply (smol, 2008). one of the best and most powerful biological proxies which can be studied in lake sediments for the reconstruction of lake trophic evolution, both related to climate and humans, is represented by cladocera remains (zawisza and szeroczyńska, 2007; chen et al., 2010). cladocera, as a key component of zooplankton, represent a biotic component of the lake ecosystem and provide important information on the lake environment (lampert and sommer, 2007). on the other hand, pollens represent one of the best and most important sources of information about changes in terrestrial vegetation and land use. the interpretation of pollen data offers the potential for reconstructing human impact on the lake catchment in prehistoric times (behre, 1986; birks et al., 1988). thus, the combination of subfossil cladocera studies and palynological analyses can enhance the reliability of reconstruction of past water trophic level and land-use within the lake catchment. in this study, we present the reconstruction of changes in trophic status of lake jelonek (n poland) from its origin until modern times, with the objective of assessing the relative contribution of natural and anthropogenic sources of nutrients on the basis of subfossil cladocera and pollen analysis. the long sediments record of lake jelonek, which covers a time span of ~14,000 years, offered a good opportunity to understand the relative contribution of climate and human activity to the development and dynamics of aquatic ecosystems in this region. from the late glacial to the early holocene the record provided insights into the dynamics of a system driven mainly by natural forces. contrastingly, the study of the regional and local conditions from the mid holocene to modern times provided scenarios for different human influences on the landscape. thus, this time period mostly contributed to understanding lake dynamics under contrasting technological systems of land use. pollen indicator taxa provided information on the main modes of human land use of the lake catchment, while basic geochemistry and cladocera provided insights into the impact of human activities on the trophic status of the lake. methods study site lake jelonek is located in central northern poland, in the south-eastern part of tuchola forest (fig. 1 a,b), at an altitude of 91 m asl the lake is of glacial origin and located in a nw-se oriented glacial channel on the ice line of the last glacial maximum close to the wda river (błaszkiewicz, 2006). lake jelonek is a small shallow water body, with an area of 20.9 ha and an average depth of 4.5 m (zmax 13.8 m, fig. 1c). the length of the shoreline is 2250 m and the shoreline development ratio is small (1.39). the water ph is circa neutral (~7.2), and the conductivity is around 300 µs cm–1. the modern trophic status of lake jelonek has been determined as mesotrophic (jańczak, 1997). the tuchola forest region is characterized by a climate from the group d – continental/microthermal climates (mcknight and darrel, 2000; przybylak et al., 2010; woś, 2010). mean annual air temperature and precipitation are 7.6°c and 558 mm, respectively (filbrandt-czaja, 2009). the winter lasts around 70 days, with a mean temperature of -3.2°c. the duration of ice-cover is approximately 3 months, from the beginning of january until the end of march, with a growing season of approximately 210 days per year. the modern vegetation is dominated by the monoculture of pine, arising from a one-way human activity. nowadays, area the natural forest ecosystems are almost absent in the tuchola forest (boiński, 1992). sampling and analyses in winter 2002, a 1330 cm long sediment core was recovered from lake jelonek using a livingstone-type corer. the sampling point (53°46.1’ n, 18°23.5’ e) was located in the deepest part of the lake (water depth ~13.8 m, fig. 1c). after coring, the sediment was packed and transported to the laboratory for a lithological description, and a subsampling procedure, which was carried out at 5 to 10 cm intervals. subsamples were analysed for subfossil cladocera, pollen, and basic geochemistry. fifteen points along the core were sampled for the radiocarbon dating of the sedimentary sequence. twelve terrestrial plant macrofossils samples were 14c dated by ams (accelerator mass spectrometry), whereas three of bulk sediment samples were analyzed by 14c conventional methods (stuiver and polach, 1977). all radiocarbon analyses were carried out at gadam absolute dating methods centre, gliwice, poland (tab. 1). ages were calibrated according to the intcal 13 curve (reimer et al., 2013). selected pairs of depths and calibrated ages were used to fit a bayesian age-depth model using bacon (blaauw and christen, 2011). carbonate and organic matter content were measured in 76 sediment subsamples (10 cm3), which were firstly dried at 105°c and homogenized in an agate mill, and then subjected to a two-step procedure (bengtsson and enella, 1986). in the first reaction, organic matter was oxidized at 550°c to carbon dioxide and ash. in the second reaction, carbon dioxide was evolved from carbonate at 925°c, leaving oxide. the carbonate content was calculated using the relationship carb =1.36 loi925, where no n c om me rci al us e o nly millenial-scale trophic changes of lake jelonek 165 fig. 1. location of lake jelonek. a) location in europe. b) location of lake jelonek (asterisk) in the tuchola forest area, boundaries of tuchola forest area according to kowalewski (2002). c) bathymetric map of lake jelonek with the coring point marked at the lake deepest point. no n c om me rci al us e o nly 166 e. zawisza et al. loi925=co2 evolved from the sample, as defined by heiri et al. (2001). the subfossil cladocera analysis was conducted on 143 sediment subsamples (1 cm3), which were processed according to the standard method proposed by frey (1986). samples were boiled in a 10% solution of koh, and later left for half an hour in a magnetic stirrer to eliminate organic matter. the residue was washed with distilled water and sieved through a 35 μm mesh. the final residue was dissolved in 10 ml–1 distilled water. for every microscope slide, 0.1 ml of final solution was used and the identification of remains was made at magnifications of 100x, 200x, and 400x, using an olympus bx40 transmitted-light microscope. two to four slides (minimum of 200 remains) from each sample were counted. all cladoceran remains were counted (head shields, shells, postabdomens, ephippia), and complete individuals were added up from body parts. cladocera identification was based on szeroczyńska and sarmaja-korjonen (2007) and flössner (2000). palynological analyses were conducted on 137 sediment subsamples (1 cm3) accounting at a resolution of 510 cm. samples were treated to eliminate non-pollen components of the sediment according to standard protocols (berglund and ralska-jasiewiczowa, 1986), and acetolysis was conducted to remove cellulose. two lycopodium tablets were added to each sample prior to acetolysis in order to allow the calculation of absolute concentrations of sporomorphs (stockmarr, 1971). treated samples were stained with fuchsine and immersed in glycerine, and sporomorphs were counted under a microscope zeiss at 340 and 680 magnifications. at least 1000 pollen grains were counted from each sample, with total pollen counts being lower only in bottom samples, because of very low pollen amounts in the deepest core section. identification of pollen grains and spores was based on erdtman et al. (1961), faegri and iversen (1989), reille (1992), and beug (2004). regional pollen assemblages reported as markers of chronostratigraphic units for the studied time period were used to refine the chronology after filbrandt-czaja (2009). all arboreal (ap) and herbaceous (nap) taxa were included in the pollen sum that served as the basis to calculate pollen percentages. simplified pollen diagrams were constructed using the c2 software (juggins, 2007) to summarize the changes of selected taxa, mainly trees and herbaceous plants considered as indictors of human impact (behre, 1981). the identified taxa were subdivided into six major ecological groups (after berglund 1969): i) trees preferring damp soils, e.g. alnus; ii) highly competitive and shade-tolerant trees; iii) trees immigrating during late holocene, e.g. fagus, carpinus, picea; iv) scarcely competitive and light-demanding trees, often belonging to the primary stage of holocene forest successions; v) apophytes, i.e. shrubs and herbs favoured by man; vi) anthropochors, i.e. herbs and graminoids introduced by man. core depth are indicated in all stratigraphic plots as cm below the lake surface (bls), in order to allow comparison with previous paleolimnological studied on lake jelonek (filbrant-czaja 2009; błaszkiewicz, 2006). results sediment lithology and geochemistry the sediments of lake jelonek were mainly composed of detritus and calcareous gyttja (fig. 2). below 2705 cm bls, the sediment consisted of light beige sand. between 2705 and 2695 cm bls, the sediment was dominated by sandy silt (ca. 20% caco3 content, fig. 2), topped by a 5 cm-thick layer of sand (2695-2690 cm). from 2690 to 2580 cm bls, the sediment was composed of light grey silt with an increasing carbonate content reaching up to 40%. from 2580 to 2500 cm bls, sediment was characterized by dark olive-beige silt detritus gyttja (organic matter ca. 50%, caco3 >10%). from 2500 to 2400 cm bls, light grey detritus gyttja with a very low content of caco3 >5% was identified. between 2400 and 2160 cm bls calcareous-detritus gyttja with occasional laminations occurred. dark detritus gyttja with a high content of organic matter and low caco3 dominated from 2160 to 1800 cm bls. from 1800 to 1530 cm bls, sediments were composed of dark calcareous-detritus gyttja with occasional lamination. from 1530 to 1480 cm bls, the sediment was dark detritus gyttja, whereas the uppermost part of the sedimentary deposit (1480 to1380 cm bls) consisted of dark brown-olive, calcareous detritus gyttja, with high caco3 content (on average ca. 45%). chronology the age-depth model was constructed using 12 radiocarbon dates (tab. 1) and 5 ages defined by the appearance of biostratigraphic markers of the central europe chronostratigraphy (mangerud et al., 1974; walanus and nalepka, 2010). three pollen-based dates were rejected due to stratigraphic inconsistencies. thus, a total of 17 depth-age data pairs were used to fit the bayesian agedepth model (fig. 3), yielding a basal age of ~14,000 calibrated years before present (cal yr bp) for the whole sedimentary sequence. subfossil cladocera a total of 29 cladocera species belonging to five families were identified in the sedimentary sequence of lake jelonek. planktonic species were represented by the families bosminidae, daphniidae, and leptodoridae, as well as by littoral species of chydoridae and sididae. planktonic species accounted to more than 60% of total cladono n c om me rci al us e o nly millenial-scale trophic changes of lake jelonek 167 cera relative abundance, thus representing the major component of the cladocera assemblages during the last 14,000 years (fig. 4). six cladocera assemblages zones (caz) were empirically distinguished based on the species composition and changes in the relative abundance of the species (fig. 4). caz i, before 12900 cal yr bp (2705 – 2640 cm bls) the initial phase of lake jelonek development was characterized by low cladocera abundance and low species richness. eight cladocera species were identified, with a maximum concentration of 800 specimens per cm3 of sediment (fig. 4). pelagic species, such as bosmina longirostris (o.f. müller, 1757) and daphnia longispina group (o.f. müller, 1785), were dominant (with a relative abundance over 80%). littoral species tolerant to cold waters, i.e. alona affinis (leydig, 1860), chydorus sphaericus (o.f. müller, 1785), acroperus harpae (baird, 1835), alonella nana (baird, 1843), were also present. caz ii 12900 9600 cal yr bp (2640-2540 cm bls) this phase was characterized by a sudden increase in cladocera abundance. a total of 20 species were identified, consisting mainly in planktonic taxa, i.e. bosmina longirostris, bosmina (e.) coregoni (baird, 1857), daphnia pulex group (leydig, 1860), and leptodora kindti (focke, fig. 2. schematic lithology and depth profiles of organic matter, carbonate and mineral fraction content in the sediments from lake jelonek; bls, below lake surface. no n c om me rci al us e o nly 168 e. zawisza et al. 1844), which accounted for over than 60% of total cladoceran remains (fig. 4). noteworthy is the presence of species associated to warm waters, such as camptocercus rectirostris (schoedler, 1862), and pleuroxus trigonellus (o.f. müller, 1776), and of species living in association with aquatic plants and having higher edaphic requirements, such as chydorus sphaericus and species from the genera alona (baird, 1843) and alonella (sars, 1862). caz iii 9600 6200 cal yr bp (2540 2420 cm bls) cladocera species richness increased through this phase to up to 23. pelagic bosmina longirostris, bosmina (e.) coregoni, and bosmina (e.) longispina dominated the cladoceran community. the latter species showed its first appearance and reached its maximum abundance through the entire sediment record within this zone. a total of 17 littoral species were identified in caz iii, the majority being reported as occurring in association with aquatic plants, such as alona and alonella spp. as well as acroperus harpae, pleuroxus sp. and sida crystallina (straus, 1820). three species showed their first appearance within this phase, namely alonella excisa (fischer, 1854), leydigia leydigi (schoedler, 1858), and graptoleberis testudinaria (fischer, 1854) (fig. 4). caz iv 6200 3900 cal yr bp (2420 2350 cm bls) this period was characterized by a major change in the structure of the cladoceran community. only one pelagic species, i.e. bosmina longirostris, persisted with an average abundance of ca. 38%. littoral species dominated for the first time through the history of the lake, especially those favoured by higher trophic status of the water. the most abundant of these species were alona rectangula (sars, 1862) (35%) acroperus harpae (10%), alonella excisa (9%), and pleuroxus trigonellus (8%). caz v 3900 1500 cal yr bp (2350 – 1730 cm bls) the flourishment of both pelagic and littoral cladoceran species characterized this period. the upper and lower boundaries of this phase were marked by an increase in the abundance of alona rectangula and chydorus sphaericus. these two species together with bosmina longirostris accounted for over 75% of all cladocerans (fig. 4). the early stages of caz v was marked by the disappearance of pelagic species associated to lower trophic status waters, i.e. bosmina (e.) longispina and bosmina (e.) coregoni, and by the decline of alona rectangula and chydorus sphaericus. in the second half of this phase, the abundance of planktonic species increased again: bosmina (e.) coregoni and bosmina (e.) longispina reached 17% and 11%, respectively, of the whole cladoceran remains. caz vi 1500 cal yr bp – present (1730 1380 cm bls) this phase was characterized by a clear decline in the abundance of cladocera, namely from an average of 27000 ind. cm–3 in phase v down to 7000 ind. cm–3 in phase vi. pelagic species (bosminidae) dominated and tab. 1. 14c dates from lake jelonek sediments. sample name depth type of laboratory number 14c yr bp cal yr bp (cm bls) dating jelonek-1-1450* 1450 ams gda-823 172±30 340-302 (17.4%) 276-185 (50.9%) jelonek-2-1550* 1550 ams gda-824 417±30 560-484 (86.7%) jelonek-3-1660* 1660 ams gda-825 1382±35 1405-1311 (95.2%) jelonek-4-2060* 2060 ams gda-826 1742±35 1782-1606 (95.4%) jelonek-5-2120° 2120 ams gda-827 1647±35 1496-1596 (68%) jelonek-6-2260° 2260 ams gda-828 2222±35 2181-2307 (68%) jelonek-7-2350* 2350 ams gda-829 3462±35 3883-3691 (95.4%) jelonek-8-2380* 2380 ams gda-830 4222±40 4909-4839 (38.5%) 4895-4676 (56.9%) jelonek-10-2480* 2480 conv. gd-30100 6702±180 7922-7322 (93.3%) jelonek-11-2540* 2540 conv. gd-30098 7852±120 9048-8578 (95.4%) jelonek-12-2560° 2560 conv. gd-15878 8602±140 9499-9857 (68%) jelonek-13-2580* 2580 ams gda-832 10252±50 12219-10814 (94.1%) jelonek-14-2600* 2600 ams gda-833 10882±60 12948-12737 (95.4%) jelonek-15-2660* 2660 ams gda-834 10832±60 12872-12713 (95.4%) jelonek-17-2705* 2705 ams gda-836 13192±70 13221-15693 (95.4%) bls, below lake surface; *samples included in the age-depth model; °samples excluded from the age-depth model because of stratigraphical inconsistencies; ams, acceleration mass spectrometry; conv., conventional 14c dating. no n c om me rci al us e o nly millenial-scale trophic changes of lake jelonek 169 accounted for about 60% of total cladoceran abundance. the contribution of alona rectangula and chydorus sphaericus significantly declined, accounting only for 8% and 6% of the total abundance, respectively (fig. 4). pollen palynological taxa that are considered as indicators of water fertility, e.g. nymphaea (l.), nuphar (sm.) or potamogeton (l.) occurred at small percentages (<2%) throughout the entire stratigraphic profile, and therefore could not be used to reconstruct the trophic evolution of lake jelonek (filbrandt-czaja, 2009). based on the analysis of pollen taxa which are considered as indicators of human activity four so-called human colonization phases were distinguished (fig. 5). pollen showed first evidence of human activity in the basin of lake jelonek during the atlantic period, which identifies the first settlement phase (ca. 7.8 3.85 cal kyr bp between 2470 and 2350 cm bls). the phase covered the period of mesolithic and neolithic economic recovery and is characterized by a decline of elm (by 6%), oak (by 5%) and ash (by 2%), along with an increased abundance of pine (fig. 5). the presence in this core section spores of fern (pteridium aquilinum (l.) kuhn), a fern species occurring on soils containing ash (tinner et al., 2000), confirms the presence of human settlements in the region. at the end of this phase, the first pollen grains of ribwort plantain (plantago lanceolata (l.)) were identified. however, indicators of anthropic pressure occurred with a relatively low percentage in this core section. the beginning of the second settlement phase (3.6 2.9 cal kyr bp, 2300-2140 cm bls) coincided with the bronze age. this period was represented in the pollen diagram (fig. 5) by deforestation, in particular a decline of hornbeam (carpinus betulus (l.)) as well as oak (quercus (l.)), elm (ulmus (l.)), linden (tilia (l.)) and hazel (corylus avellana (l.)). reduction in the number of mesophilous deciduous trees was probably connected with the objective of settlers to acquire fertile lands overgrown with forests. this period was also characterized by an increased content of mugwort pollen (artemisia (l.)), a ruderal plant which indicates presence of crops as it was a common weed of primitive cereal crops (behre, 1981). the increase in the content of heather (calluna vulgaris (l.) hull) pollen along with the presence of common jufig. 3. age-depth model based on the 14c dating of lake jelonek sediments, showing sediment accumulation from late glacial to current times. stars indicate 14c date (see tab. 1), and arrows point the chronostratigraphical (pollen) borders according to mangerud et al. (1974) and walanus and nalepka (2010). no n c om me rci al us e o nly 170 e. zawisza et al. niper (juniperus communis (l.)) and broadleaf/hoary plantain (plantago (l.) major/media) pollen indicate the presence of canopy gaps caused by cattle grazing, as well as an increase in the area of dry pastures. the end of the second settlement phase was characterized by a substantial increase in the abundances of hornbeam (fig. 5). the third settlement stage encompasses the iron age (ca. 2.8 1.7 cal kyr bp, 2110-1790 cm bls) and was distinguished by a significant increase of palynological anthropogenic indicators and by the decline of deciduous trees such as hornbeam, linden, ash and elm (fig. 5). the most important feature of this period was the increasing abundance of cereal pollen, mainly rye (secale cereale (l.)), and sorrel pollen (rumex acetosa (l.)/acetosella (l.)). the upper border of this phase is marked by interruption or significant descent of percentages of all herbaceous taxa (fig. 5). the fourth and last settlement stage (1.3 cal kyr bp up to the present, 1660-1380 cm bls) began in the early middle age and was characterized by the increase of herbaceous taxa and the decline of deciduous tree species pollen (fig. 5). a number of palynological taxa occurred, which fig. 4. relative abundance, total cladocera sum (as individuals cm–3 of fresh sediment), and planktonic/littoral ratio of sediment subfossil cladocera of lake jelonek from late glacial to current times; bls, below lake surface. fig. 5. simplified, pollen diagram from lake jelonek. the represented taxa are subdivided into six ecological groups according to berglund (1969): i) trees on damp soils; ii) highly competitive and shade-tolerant trees; iii) trees immigrating during late holocene; v) scarcely competitive and light-demanding trees; v) shrubs and herbs favoured by man (apophytes); vi) herbs and graminoids introduced by man (anthropochors). ap, arboreal pollen; nap, non-arboreal and herbs pollen; bls, below lake surface. no n c om me rci al us e o nly millenial-scale trophic changes of lake jelonek 171 are considered as indicators of grazed forests, meadows and pastures, such as poaceae, rumex sp., plantago lanceeolata, ranunculaceae, calluna vulgaris, campanulaceae. pollen grains of cultivated plants, e.g. cerealia type, secale cereale, wheat (triticum (l.)) occurred in large numbers. discussion based on the results of geochemical analyses, subfossil cladocera remains and pollens the trophic conditions in lake jelonek were reconstructed from the late glacial to modern times. changes in the composition of cladocera species, and in particular of taxa indicating higher lake trophic conditions (e.g. bosmina longirostris, chydorus sphaericus, alona rectangula), enabled the assessment of the trophic history of lake jelonek. changes in the afforestation rate (inferred by increasing or decreasing percentages of arboreal pollens), together with the presence of pollen of synanthropic and cultivated plants, helped to determine the type of economy (breeding, cultivation) and its impact on the vegetation cover of the lake catchment (berglund, 1969; aaby, 1986). lacustrine sedimentation in lake jelonek began during the late glacial. the pollen spectrum that characterized this time period (caz i, before 12900 cal yr bp) indicates that the plant communities resembled the vegetation of the arctic tundra where heliophytes and plants tolerant to low temperature and unfavourable edaphic conditions are dominant (filbrandt-czaja, 2009). the low cladocera species richness (only eight species) indicated a relatively shallow lake characterized by cold and nutrient-poor waters. the cladocera species composition and the dominance of bosmina longirostris at the base of the record suggested that species typically living in the open water zone were more aggregated to the lake shores. large areas of the lake were free of plant coverage, and nutrients were available only within the shoreline zone. the content of mineral matter deposited within caz i amounted up to 95% of wet sediments, likely as a result of high supply of clastic material from the vegetation free shores or melting dead-ice blocks. species richness and abundance of cladocerans increased at the end of late glacial, indicating that climate and environmental conditions during the second half of this period were more suitable for zooplankton development. the more favourable climate conditions were confirmed by the rising abundances of pine pollen, although pollen of numerous herbaceous taxa was still present in large numbers. high relative abundances of cyperaceae indicate well-developed vegetation in the littoral zone. the higher abundances of pelagic species (i.e. bosmina (e.) coregoni) indicated a likely deepening of the lake during the stage of progressive climate warming. average annual air temperature showed a clear increase by the early holocene (ralska-jasiewiczowa et al., 1998), which in turn induced a substantial transformation of the vegetation within the catchment of lake jelonek. the first stage of succession was the growth of birch and birch-pine forests. the warming was also reflected in the succession of cladocera species, which showed a sudden and clear increase in species diversity and abundance. through the second half of caz ii (11400 9600 cal yr bp), in addition to planktonic species (family bosminidae), pelagic species leptodora kindti and daphnia pulex group appeared (szeroczyńska, 1993; korhola, 1999; amsink et al., 2005), and accounted together to ca. 60% of the cladoceran assemblage. a substantial increase of planktonic species suggested high water levels during the preboreal period. an increase of species that live in association with aquatic plants was also observed, in particular of species belonging to the family alonidae, and species associated with warmer and clearer waters, such as pleuroxus trigonellus and camptocercus rectrirostris (korhola, 1999; amsink et al., 2005; szeroczyńska and zawisza, 2011). the presence of these species indicates the expansion of the littoral zone and the development of littoral vegetation. the content of organic matter in the sediment showed a gradual increase since the beginning of the holocene, and reached ~50% at the end of the preboreal period. however, despite the prevalence of more favourable climatic and edaphic conditions during the early holocene, the subfossil cladocera indicate no significant changes in lake trophic status, which probably remained at the oligo-mesotrophic level (fig. 6). this persistence of low nutrient levels in the water might have been possibly related to dilution by the rising of water level. the development of forests, which initiated during the preboreal period, continued through the boreal one, and was mainly characterise by increasing pine accompanied by mesophilous trees species, namely corylus avellana, alnus, and ulmus. this period was also characterized by an increased frequency and abundance of cladocera species (at the beginning of caz iii, 9600-8400 cal yr bp). bosmina (e.) longispina, a species preferring welloxygenated waters with low trophic status, occurred in the pelagic zone and reached its maximum abundance through this zone. the presence of both pelagic and littoral taxa indicates that the lake was fairly deep and its littoral zone was well developed. high values of organic matter (over 50%) deposited in the sediments during this period confirmed the richness of life in the littoral zone and an oligo-mesotrophic status (fig. 6). during the holocene climate optimum the vegetation around lake jelonek was characterized by mixed deciduous forests dominated by ulmus, quercus, tilia and fraxinus. the presence of pollen grains of ivy (hedera helix (l.)) and mistletoe (viscum (l.)) clearly indino n c om me rci al us e o nly 172 e. zawisza et al. cates mild winters and warm summers (iversen 1944, troels-smith 1960). the atlantic period (caz iii and caz iv, 8400-5800 cal yr bp) manifested in the cladocera community by the dominance of species with higher edaphic requirements (bosmina longirostris), and of species that usually occur in association with aquatic plants (e.g. pleuroxus trigonellus and acroperus harpae, both belonging to the family chydoridae). such a cladoceran assemblage indicates a probably increase in lake trophic conditions up to mesotrophic levels (fig. 6). it is possible that the increase in the trophic status of lake jelonek was the result of the first stage of human settlement, which was discriminated on the basis of palynological evidences at the end of the atlantic period. in fact, this period corresponds to the mesolithic settlement, which, however, was characterized by little groups of settlers that generally had little impact on the existing water bodies (filbrandt-czaja, 2009). the economic activity of mesolithic tribes of hunters mainly focused on forested areas where they cleared the undergrowth with fire, aiming of chasing out the game and intensify the growth of herbaceous vegetation that served as food for deer (latałowa, 1994, 2003; filbrandt-czaja, 2009). it seems likely, therefore, that the increased trophic status of lake jelonek during the atlantic period was mainly related to natural processes associated with favourable climate conditions (zawisza and szeroczyńska, 2007) rather than to human activity. since the beginning of the subboreal period, the neolithic human impact was much more evident in the sedimentary record of lake jelonek, even though transformation of the natural environment by neolithic groups of settlers still remained small, as the main subsistence strategy was still based on hunting and gathering (latałowa, 2003, 2007; filbrandt-czaja, 2009). nevertheless, human activity within the lake catchment was reflected by the succession of subfossil cladocera, by a higher abundance of species that prefer nutrient rich water (e.g. alona rectangula), and by the disappearance of pelagic species that prefer waters with a lower trophic level (bosmina (e.) longispina and bosmina (e.) coregoni). during this period, the lake was probably in mesotrophic conditions (fig. 6). fig. 6. comparison of relative abundances of cladocera species indicator of lake trophic status (i.e. bosmina longirostris, alona rectangula, chydorus sphaericus) and total cladocera abundance sum with settlement phases identified by pollen analyses. trophic reconstruction based on subfossil cladocera analysis. caz, cladocera assemblages zones; bls, below lake surface. no n c om me rci al us e o nly millenial-scale trophic changes of lake jelonek 173 between 3850-3600 cal yr bp , cladocera species preferring low trophic conditions, such as bosmina (e.) coregoni and bosmina (e.) longispina (hofmann, 1977; korhola, 1999; brancelj et al., 2009; zawisza et al., 2016) further decreased, while alona rectangula declined. such changes in the zooplankton community indicate a recession in the trophic status of lake jelonek back to oligomesotrophic levels, which was most likely related to a gap in the human colonization process (between settlement phase 1 and 2), which was inferred based on the pollen evidence (fig. 6). through the middle and late subboreal period, there were substantial changes in the cladocerans succession (caz v). whereas alona rectangula and chydorus sphaericus increased during this period (reaching the highest values in the profile), bosmina (e.) coregoni significantly decreased, and bosmina (e.) longispina disappeared from the record. such species composition suggests that the lake was affected by an eutrophication process at the time, probably caused by an increased supply of nutrients into the water. the increasing curve of species considered indicators of increased trophic status (chydorus sphaericus, alona rectangula) was perfectly synchronous with the second and third settlement stages. i.e. during the bronze and iron age, respectively, which both were determined based on palynological analysis (fig. 6). the occurrence of grasses (poaceae), plantago lanceolata, plantago major/media, rumex sp. suggests that livestock played an important role in the human economy. most likely they were grazed in thinned forests where large areas were covered with heather, cow wheat (melampyrum (l.)), and bracken. these activities played a major role during the bronze and iron age (filbrandtczaja, 2009), and continued during the roman influence (wielowiejski, 1981). during this time humans likely settled close or around lake jelonek, so that cattle breeding and other human activities resulted in an increased supply of nutrients to the lake, which increased the lake productivity up to eutrophic conditions through caz v. the increase in the organic matter content in the sediments layers of lake jelonek deposited during this period offers additional support to the interpretation of increased lake trophic status. it is important to mention that significant changes in pollen spectra were recorded toward the end of the second settlement, especially consisting in an increasing abundances of hornbeam trees. this was high probably related to the colonization of abandoned fields and grazing lands by this arboreal taxon, which suggests that settlers abandoned the environs of lake jelonek. the recovery of the forest cover marked a discontinuity in the colonization, which lasted for around 150 years between phase 2 and 3. the second half of the subatlantic period, corresponding to the migration period at the beginning of caz vi, was characterized by a significant decrease in the abundance of cladocera species. in particular, the frequency of species indicating high trophic status declined (i.e. alona rectangula and chydorus sphaericus), whereas the contribution of pelagic species bosmina (e.) longispina and bosmina (e.) coregoni increased. such a species assemblage indicate a reduction in the trophic status of lake jelonek down to mesotrophic conditions, which persisted till present day. the improvement of the lake water quality was probably associated with human populations abandoning the lake jelonek basin. this process began about 1600 years ago in the period referred to as “pre-modern human migration” (guy, 2007; kobyliński, 2015). significant regression of colonization between iv-vi century ad and the related forest regeneration, were reflected by the recovery of forest species and by a major decline of all herbaceous plants. indicators of human occupation dropped to a minimum, which outline an encroachment of broadleaf forest dominated by hornbeam on areas abandoned by man. the fourth settlement phase (fig. 6) began during the early middle age and continued to present day. the pollen sequence from lake jelonek indicates that human activity through the early middle age was accompanied by land deforestation, probably associated with land reclamation for grazing and cereal cultivation. the high content of cereal pollen clearly indicates an increased intensity of colonization compared to the previous period, as well as the dominant role of cereal crops in the human diet. however, according to archaeological research, the region of lake jelonek was sparsely populated during the middle age (wawrzeniuk, 2005). probably, temporal human settlements characterized this time period (woźny, 2005), which may imply a significant decrease in the amount of biogenic substances delivered to the lake. indeed, a considerable decline of alona rectangula and chydorus sphaericus, both indicators of high trophic status, offer further support for the inference of small human populations. historical data suggests that the impact of human colonization in the vicinity of lake jelonek gradually increased since the 15th century (chudziak, 2000). maps from the late 18th century (schrötter maps) show that at that time the lake was surrounded to the north and northeast by arable fields and meadows, as well as by pine forests to the south and west. the land-use change affected the structure of cladocera zooplankton, which consisted in the decrease of species usually associated with oligotrophic clear waters, such as bosmina (eubosmina) and alona affinis. the most extensive deforestation in the region of lake jelonek occurred in the first half of the 19th century (ślaski, 1951), i.e. in the period associated with the uppermost sediments layer (caz vi). abundance of species no n c om me rci al us e o nly 174 e. zawisza et al. that usually occur in nutrient rich waters (bosmina longirostris and alona rectangula) increased at that time. since the early 20th century the lake has been surrounded again by a dense forest (broda, 2000). afforestation represent the major land use change since then, and has resulted in a considerable isolation of the lake from the impact of human economic activity. after the lands around lake jelonek were abandoned by the local population, the nutrient input to the lake decreased considerably. this allowed the beginning of a process of so-called self-purification of the lake, which finally returned to present mesotrophic conditions. conclusions the environmental and ecological development of lake jelonek through the late glacial and the early holocene appeared to be mainly driven by climate changes, whereas the impact human activities became prevailing over climate related factors through the mid and late holocene. except during its initial formative phase, lake jelonek permanently was a relatively deep water body, while its trophic status oscillated several times between oligoand eutrophic level. oligotrophic conditions prevailed from the late glacial to the early holocene, while eutrophic conditions dominated during the settlement stages of the bronze and iron age. during the remaining periods the lake mesotrophic conditions prevailed. given the lake size and depth and the population density near the shores, it can be concluded that the increase in the trophic status during the subboreal and subatlantic period was primarily the results of human economic activity. the last century has been characterized by the extensive recovery of the forest cover, which has substantially contributed to the recovery of a mesotrophic lake status. the results presented here provide new important information on effects of climatic changes on the tuchola forest area, and on the edaphic condition of its lakes during late glacial and holocene. the data presented, especially those on subfossil cladocera analysis, integrated other paleolimnological studies, thus contributing at reducing the existing knowledge gap regarding causes and mechanisms of water eutrophication during the last millennia. the present study also stressed the close relationship that exists between lake water quality and lake catchment processes. changes in forest cover connected with human activity revealed to have played the major role in driving lake trophic changes, which was not outlined for the area of tuchola forest before. the results of the present study can contribute to a better understanding of effects of climate change at anthropogenic influence on long term trophic fluctuation of lakes in northern poland and central europe. acknowledgments this research was possible with the support of the institute of geological sciences, polish academy of sciences (ing pan), polish ministry of science (n n306 228039) and by the polish german governments (mnisw-daad, 2016-2017) daad project. references aaby b, 1986. trees as anthropogenic indicators in regional pollen diagrams from eastern denmark, p. 73-94. in: k.e. behre (ed.), anthropogenic indicators in pollen diagrams. a.a. balkema, rotterdam. amsinck sl, jeppesen e, landkildehus f, 2005. relationships between environmental variables and zooplankton subfossils in the surface sediments of 36 shallow coastal brackish lakes with special emphasis on the role of fish. j. paleolimnol. 33:39-51. berglund be, 1969. vegetation and human influence in south scandinavia during prehistoric time, 12:9-28. in: b.e. berglund (ed.), impact of man on the scandinavian landscape during the late post-glacial. oikos suppl. 12:9-28. berglund be, ralska-jasiewiczowa m, 1986. pollen analysis and pollen diagrams, p 455-484. in: b.e. berglund (ed.), handbook of holocene palaeoecology and palaeohydrology. j. wiley & sons ltd., chichester. behre ke, 1981. the interpretation of anthropogenic indicators in pollen diagrams. pollen et spores 23:225-245. behre ke, 1986. anthropogenic indicators in pollen diagrams. a.a. balkema, rotterdam: 232 pp. behre ke, 1992. the history of rye cultivation in europe. veg. hist. archaeobot. 1:141-156. bengtsson l, enell m, 1986. chemical analysis, p. 423-45. in: b.e. berglund (eds.), handbook of holocene palaeoecology and palaeohydrology. j. wiley & sons ltd., chichester: 869 pp. beug hj, 2004. [leitfaden der pollenbestimmung für mitteleuropa und angrenzende gebiete].[book in german]. verlag dr. friedrich pfeil, münchen: 542 pp. birks hh, birks hjb, kaland pe, 1988. the cultural landscape past, present and future. cambridge university press, cambridge: 521 pp. blaauw m, christen ja, 2011. flexible paleoclimate age-depth models using an autoregressive gamma process. bayesian analysis 6:457-474. błaszkiewicz m, 2006. [szczegółowa mapa geologiczna polski w skali 1:50 000 wraz z objaśnieniami – arkusz osiek (167)].[map in polish]. państwowy instytut geologiczny. boiński m, 1992. [osobliwości szaty roślinnej borów tucholskich (przewodnik)].[book in polish]. towarzystwo miłośników borów tucholskich, toruń. brancelj a, kernan m, jeppesen e, rautio m, manca m, šiško m, alonso m, stuchlík e, 2009. cladocera remains from the sediment of remote cold lakes: a study of 294 lakes across europe. adv. limnol. 62:269-294. broda j, 2000. [historia leśnictwa w polsce].[book in polish]. wydawnictwo akademii rolniczej im. augusta cieszkowskiego w poznaniu, poznań: 368 pp. no n c om me rci al us e o nly millenial-scale trophic changes of lake jelonek 175 chen g, dalton c, taylor d, 2010. cladocera as indicators of trophic state in irish lakes. j paleol. 44:465-481. chudziak w, 2000. [badania archeologiczne], p. 23-30. in: w. chudziak (ed.), [mezoregion osadniczy warlubie-płochocinek: krajobraz przyrodniczy i kulturowy].[book in polish]. fundacji na rzecz nauki polskiej. erdtman g, berglund be, praglowski j, nilsson s, 1961. an introduction to a scandinavian pollen flora. almqvist and wiksell, stockholm: 89 pp. faegri k, iversen j, 1989. textbook of pollen analysis, 4th ed. the blackburn press, caldwell: 328 pp. filbrandt-czaja a, 2009. [studia nad historią szaty roślinnej i krajobrazu borów tucholskich].[book in polish]. wydawnictwo naukowe uniwersytetu mikołaja kopernika, toruń: 131 pp. flössner d, 2000. [die haplopoda und cladocera (ohne bosminidae) mitteleuropas].[book in german]. backhuys publishers, leiden: 428 pp. frey dg, 1986. cladocera analysis, p. 667-692. in: b.e. berglund (eds.), handbook of holocene palaeoecology and palaeohydrology. j. wiley & sons ltd, chichester. gałka m, tobolski k, zawisza e, goslar t, 2014. postglacial history of vegetation, human activity, and lake-level changes at lake linówek in northeast poland, based on multiproxy data. veg. hist. archaeobot. 23:123-152. goransson h, 1986. man and forests of nemoral broad-leafed trees during the stone age. striae, 24:143-152. grabowska j, zawisza e, jaskółka a, obremska m, 2015. natural transformation of the romoty paleolake (ne poland) during the late glacial and holocene based on isotopic, pollen, cladoceran and geochemical data. quatern. int. 386:171-185. guy h, 2007. barbarian migrations and the roman west, 376568. cambridge university press, cambridge: 614 pp. heiri o, lotter af, lemcke g, 2001. loss on ignition as a method for estimating organic and carbonate content in sediments: reproducibility and comparability of results. j. paleolimnol. 25:101-110. hofmann w, 1977. bosmina (eubosmina) population of the grosser segeberger see during late-glacial and postglacial times. arch. hydrobiol. 80:349-359. iversen j, 1944. viscum, hedera and ilex as climate indicators. a contribution to the study of the post-glacial temperature climate. geologisk föreningens förhaandlinger 66:463-483. jańczak j, 1997. [atlas jezior polski].[in polish]. instytut meteorologii i gospodarki wodnej. bogucki wydawnictwo naukowe. joye sb, smith vh, howarth r, 2006. eutrophication of freshwater and marine ecosystems. limnol. oceanogr. 51:351-800. juggins s, 2007. user guide c2 software for ecological and palaeoecological data analysis and visualisation user guide ver. 1.5. newcastle, university of newcastle. kobyliński z, 2015. the slavs, p. 530-537. in: p. fouracre (ed.), the new cambridge medieval history: 1008 pp. kowalewski g, 2002. [granice borów tucholskich], p. 121138. in: j. banaszak and k. tobolski (eds.) [park narodowy “bory tucholskie” na tle projektowanego rezerwatu biosfery].[book in polish]. park narodowy “bory tucholskie”, wydaw. homoni. korhola a, 1999. distribution patterns of cladocera in subarctic fennoscandian lakes and their potential in environmental reconstruction. ecography 22:357-373. lampert w, sommer u, 2007. limnoecology: the ecology of lakes and streams, 2nd ed. oxford university press, oxford: 324 pp. latałowa m, 1994. [gospodarka mezolityczna i początki rolnictwa na obszarze polskiego pobrzeża bałtyku w świetle badań palinologicznych].[article in polish]. pol. bot. se. kraków 11:135-153. latałowa m, 2003. [holocen], p. 273-308. in: k. sadowska, s. dybova (eds.), [palinologia].[book in polish]. wyd. instyt. bot. pan, kraków. latałowa m, 2007. [gospodarka człowieka w diagramach pyłkowych], p. 171-187. in: m. makohonienko, d. makowiecki, z. kurnatowska (eds.), [studia interdyscyplinarne nad środowiskiem i kulturą w polsce].[book in polish]. bogucki wydawnictwo naukowe, poznań. mangerud uj, anderson s, berglund be, donner j, 1974. quaternary stratygraphy of norden, a proposal of terminology and classification. boreas 3:109-128. mcknight tl, darrel h, 2000. climate zones and types. physical geography: a landscape appreciation. prentice hall, upper saddle river: 132 pp. nevalainen l, luoto tp, kultti s, sarmaja-korjonen k, 2013. spatio-temporal distribution of sedimentary cladocera (crustacea: branchiopoda) in relation to climate. j. biogeogr. 40:1548-1559. o’sullivan pe, 1983. annually laminated lake sediments and the study of quaternary environmental changes. quat. sci. rev.1:245-313. przybylak r, majorowicz j, brázdil r, kejna m, 2010. the polish climate in the european context: an historical overview. springer, new york: 535 pp. ralska-jasiewiczowa m, goslar t, różański k, wacnik a, czernik j, chróst l, 2003. very fast environmental changes at the pleistocene/holocene boundary, recorded in laminated sediments of lake gościąż, poland. palaeogeogr. palaeoclimatol. palaeoecol. 193:225-247. ralska-jasiewiczowa m, goslar t, madeyska y, starkel l, 1998. lake gościąż, central poland. a monographic study. szafer institute of botany, kraków. r core team, 2014. r: a language and environment for statistical computing. r foundation for statistical computing, vienna, austria. available from: www.r-project.org reille m, 1992. [pollen et spores d’europe et d’afrique du nord].[book in french]. laboratoire de botanique historique et palynologie, marseille: 520 pp. reimer pj, bard e, bayliss a, beck jw, blackwell pg, ramsey cb, buck ce, cheng h, edwards rl, friedrich m, 2013. intcal13 and marine13 radiocarbon age calibration curves 0-50,000 years cal bp. radiocarbon 55:1869-1887. richerson pj, suchanek th, zierenberg ra, osleger da, heyvaert ac, slotton dg, eagles-smith ca, vaughn ce, 2008. anthropogenic stressors and changes in the clear lake ecosystem as recorded in sediment cores. ecol. applic.18: 257-283. schmidt r, muller j, drescher-schneider r, krisai r, szeroczyńska k, barić k, 2000. changes in lake level and trophy at lake vrana, a large karstic lake on the island of cres (croatia), with respect to palaeoclimate and anthropogenic no n c om me rci al us e o nly 176 e. zawisza et al. impacts during the last approx. 16,000 years. j. limnol. 59:113-130. smith vh, schindler dw, 2009. eutrophication science: where do we go from here? trends ecol. evol. 24:201-207. smol jp, 2008. pollution of lakes and rivers: a paleoenvironmental perspective. blackwell publishing ltd., oxford: 383 pp. stockmarr j, 1971. tablets with spores used in absolute pollen analysis. pollen et spores 13:615-621. stuiver m, polach g, 1977. reporting of 14c data. radiocarbon 19:355-363. szeroczyńska k, 1993. [analiza cladocera w spągowych osadach jeziora gościąż].[article in polish]. pol. bot. se. 8:173-179. szeroczyńska k, sarmaja-korjonen k, 2007. atlas of subfossil cladocera from central and northern europe. friends of the lower vistula society: 84 pp. szeroczyńska k, zawisza e, 2011. subfossil faunal and floral remains (cladocera, pediastrum) in two northern lobelia lakes in finland. knowl. manag. manag. aquat. ec. 402:1-15. ślaski k, 1951. [zasięg lasów pomorza w ostatnim tysiącleciu].[book in polish]. instytut zachodni, poznań: 263 pp. tinner w, conedera m, gobet e, hubschmid p, wehrli m, ammann b, 2000. a palaeoecological attempt to classify fire sensivity of treesin the southern alps. holocene 10:565-574. troels-smith j, 1960. ivy, mistletoe and elm climate indicatorsfodder plants. danmarks geologiske undersøgelse: 32 pp. walanus a, nalepka d, 2010. calibration of mangeruds boundaries. radiocarbon 52:1639-1644. wawrzeniuk j, 2005. [archeologia sredniowiecza i nowożytna borów tucholskich], p. 76-97. in: w. jastrzębski, j. woźny (eds.), [dziedictwo kulturowe i przyrodnicze borów tucholskich. stan badań i potrzeby badawcze].[book in polish]. bydgoszcz-tuchola. wielowiejski j, 1981. [produkcja rolno-hodowlana], p. 315-334. in: j. wielowiejski (ed.), [prehistoria ziem polskich. późny okres lateński i okres rzymski].[book in polish]. ossolinum, wrocław. woś a, 2010. [klimat polski w drugiej połowie xx wieku]. [book in polish]. wydawnictwo naukowe uam, poznań. woźny j, 2005. [archeologia pradziejowea borów tucholskich], p. 61-75. in: w. jastrzębski, j. woźny (eds.), [dziedzictwo kulturowe i przyrodnicze borów tucholskich. stan badań i potrzeby badawcze].[book in polish]. bydgszcz-tuchola. zawisza e, szeroczyńska k, 2007. the development history of wigry lake as shown by subfossil cladocera. geochronometria 27:67-74. zawisza e, zawiska i, correa-metrio a, 2016. cladocera community composition as a function of physicochemical and morphological parameters of dystrophic lakes in ne poland. wetlands 36:1131-1142. no n c om me rci al us e o nly layout 1 introduction the sediments that accumulate year by year in natural lakes are a key source of information on past environmental and climate changes and the most valuable sediment archive are those spanning over longer time periods (cohen, 2003). sediments from contemporary temperate lake allow to track environment changes that have taken place after the end of the last glaciation (szeroczyńska and zawisza 2007, nevalainen, 2013). however, information on earlier interglacial stages is precluded. in addition, lake changes during the holocene were the result of the combination between natural processes and increasing human activities (cheddadi, 1998). the effects of natural and human driven processes is difficult to be discriminated. the analyses of sediments from ancient paleolakes, which existed and disappeared from the landscape only under the influence of natural processes, provided opportunity to study a period spanning across glaciation, the late saalian (syn. wartanian, 160-140 ka), a warming period, eemian interglacial, and another cool period, early vistulian glaciation. numerous localities with lacustrine sediments from the eemian interglacial (mis 5e) are present in europe. up to date, more than 300 sediment profiles from the eemian interglacial have been discovered in poland and analysed for pollens (granoszewski, 2003; bruj and krupiński, 2001; mamakowa, 1989; dobracka and winter 2001; kupryjanowicz, 2008, kupryjanowicz et al. 2004). the analyses of other proxies (e.g., lithology, stable isotopes, geochemistry, animal and plant remains, pigments) are still less frequent in sediments of eemian age (robertsson, 2000; mirosław-grabowska, 2005; gandouin, 2007; nazarova, 2012), though each one is able to add precious information to complete the picture of the past ecosystem. over 80 eemian interglacial sites have been discovered in north-eastern poland (kupryjanowicz, 2008). numerous eemian paleolakes of the northern podlasie indicate that during the eemian interglacial, a sort of lake district existed in this region (kupryjanowicz, 2008). nevertheless archives spanning the eemian interglacial and the vistulian glaciation are rare, especially in areas which have been influence by the vistulian ice-sheet. advances in oceanography and limnology, 2016; 7(2): 206-219 article doi: 10.4081/aiol.2016.6217 this work is licensed under a creative commons attribution-noncommercial 4.0 international license (cc by-nc 4.0). the eemian/early vistulian development of the solniki paleolake (north-eastern poland) as shown by subfossil cladocera monika magdalena niska* institute of geography and regional study, pomeranian university in słupsk, partyzantów 27, 76-200 słupsk, poland *corresponding author: monikaniska@wp.pl abstract this work presents results of a paleolimnological study focussed on subfossil cladocera analysis and on different aspects of the evolution of the solniki paleolake during the eemian/early vistulian period. the study aimed at the reconstruction of the long-term dynamics of this paleoecosystem and at defining the conditions (e.g., water level, trophic status and water temperature) of the ancient lake. paleolacustrine deposits of ca. 10 m thickness were discovered at solniki during cartographic works for the trześcianka sheet of the detailed geological map of poland. this archives recorded one full-interglacial sequence (eemian interglacial), one interstadial warming (brørup) and two stadial coolings (herning and rederstall) stages, which were confirmed by palynological analyses. the subfossil cladocera fauna from the solniki paleolake consisted in 17 species belonging to the families bosminidae, chydoridae, sididae and daphniidae. littoral species were dominant (52%), the most frequent of which were alona affinis and camptocercus rectirostris. the most abundance pelagic species were eubosmina coregoni and bosmina longirostris. the sediment species composition was quite similar to that of contemporary central european lakes. the early and the late stages of eemian interglacial were likely the most favourable periods for the cladocera development in the paleolake, in relation to higher water level, moderate water temperature and the mesotrophic state of water. a further ecologically favourable period was the brørup interstadial. the highest species richness, abundance, and diversity during the whole paleolake existence were recorded during these three periods. surprisingly, the middle of the middle eemian interglacial climate optimum appeared as an unfavourable period for the cladocera growth as it was associated with decreasing water level and pronounced climate fluctuations. this sequence was also recorded by other studies of eemian lakes in central europe. key words: cladocera analysis; eemian/early vistulian; ne poland. received: july 2016. accepted: december 2016. no n c om me rci al us e o nly eemian/vistulian development of a polish paleolake 207 solniki is one of two northern podlasie sites which play a key role in the reconstruction of vegetation succession and climate changes over the last interglacial-glacial cycle, from the final stage of the wartanian glaciation (late saalian) to the end of the early vistulian (kupryjanowicz, 2007). a paleolake sediment of ca. 10 m thickness was discovered at solniki during cartographic works for the trześcianka sheet of the detailed geological map of poland. prof. m. kupryjanowicz. started the palynological study of the solniki profile demonstrating the eemian and early vistulian age of the discovered paleolake (kupryjanowicz et al., 2005; kupryjanowicz, 2008). subfossil cladocera analysis was performed for this profile in order to complement and expand environmental information provided by the palynological data. subfossil cladocera are an useful tools in paleolimnology. they are a widespread and well studied component of the microfauna in lakes and ponds, and are commonly used as environmental indicator in aquatic ecosystem management (jeppesen et al., 2011). in fact distribution and abundance of cladocera is affected by habitat properties (korhola, 1990), as well as by factors such as trophy (szeroczyńska, 1998a, nevalainen and luoto, 2016), acidification (krause-dellin and steinberg, 1986), water temperature (poulsen, 1944; hofmann, 2000, nevalainen et al., 2013) and fish predation (jeppesen et al., 2001; korhola and rautio, 2001). subfossil cladocera analyses in poland provided detailed results for the last 13,000 years (gąsiorowski, 2002; szeroczyńska, 2006; szeroczyńska and zawisza, 2007). the capability of subfossil cladocera to support the environmental reconstruction for contemporary lakes, led to considerable interest in the applicability of the methods to sediments from the period preceding the last glaciation. literature on cladocera succession during the eemian interglacial is scarce, as only a few locations in the world have been examined, often preliminarly, for cladoceran subfossils (tsukada, 1972; kadota, 1975; hann and karrow, 1984, 1993). the first comprehensive analysis of subfossil cladocera from eemian lake sediments in europe was performed in denmark (frey, 1962). in poland, the first analyses were conducted with material obtained in konin (szeroczyńska, 1991). so far, cladocera analysis of sediment from profiles of the eemian age was performed in poland and in northern germany (börner et al., 2015) for ca. 16 sites (niska, 2008; niska, 2015; mirosław-grabowska et al, 2016, pawłowski, 2011). however, only three out of these profiles represent eemian interglacial and early vistulian sequences: solniki, kubłowo (mirosław-grabowska et al. 2016) and ustków (kołaczek et. al. 2016). very general results of the cladocera analysis from the eemian part of the solniki paleolake profile have been previously published (kupryjanowicz et. al, 2005; miroslaw-grabowska, 2015). this paper presents for the first time full results of subfossil cladocera analysis for the entire profile (eemian interglacial and early vistulian part 3.0-1.0 m depth) of the solniki paleolake sediments. the aim of this study was to reconstruct the evolution of the solniki paleolake during its long existence, and its water conditions (water level, trophic state and water temperature) based on cladoceran investigations. these results were compared to the previously published pollen data, which define the vegetation history of the surrounding area. special attention was paid to the early vistulian period, because of the scarcity of investigation of subfossil cladocera for this period. methods study site and chronology of the profile the solniki site (53°30′n, 23°12′ e; 143 m asl) is located in the central part of the białystok upland (ne poland, fig. 1), approximately 18 km south of białystok and ~4 km south-west of zabłudów (kupryjanowicz, 2008). the region’s relief was formed by the wartanian glaciation (saalian, riss glaciation) (musiał, 1992). the maximal range of the last glaciation (vistulian, weischelian) lies ~50 km north of this site (marks, 2002). the paleolake lies ~100 m east of the solniki to koźliki road, on the left bank of the river czarna. the studied sediments fill the subglacial basin, which was formed during the wartanian glaciation (kurek and preidl, 2001) and extends over an area of ~700×300 m. the paleolake is surrounded by moraine and kame hills. coring was carried out in the 1990’s using a geological corer during the preparation of the trześcianka sheet of the 1:50,000 (detailed geological map of poland, in a scale of kurek and preidl, 2001). the entire cored sequence was 162.50 m long, while lake deposits, were present at depths of between 3.00-11.00 m (kupryjanowicz, 2008). the paleolake sediments mainly consisted of organic silts, whereas a peat layer was detected at 11.0010.60 m depth below the surface, and appeared covered by a 3 m thick deposit of sand. the description of the lacustrine-mire series is shown in tab. 1. sediments from the solniki paleolake profile have been made available for cladocera analysis by courtesy of prof m. kupryjanowicz. high resolution pollen analysis was performed for the sediment section between 3.00 and 11.06 m (kupryjanowicz, 2008). the lower part of the solniki pollen profile represents the interglacial vegetation succession and its characteristic features allow for its unquestionable correlation with the eemian interglacial (kupryjanowicz, 2008). stages of vegetation opening are present in the middle part of the pinus zone (s-7 l paz), which ended the eemian succession. five local pollen zones (from sno n c om me rci al us e o nly 208 m.m niska 8 to s-12 l paz) distinguished in the upper part of the studied profile represent the early vistulian (weichselian) herning stadial, brørup interstadial sensu lato (or amersfoort/brørup interstadial sensu stricto with cold oscillation between them) and redestall stadial (kupryjanowicz, 2008). all local pollen assemblage zones determined in the solniki profile strongly correspon to the regional zones distinguished by kupryjanowicz (2008) for the northern podlasie and by mamakowa (1989) for the territory of poland (tab. 2). the solniki deposits have not been radiometrically dated because the entire eemian interglacial is outside the range of radiocarbon dating, and the carbonate lacustrine sediments are very problematic for other absolute chronometric dating methods (mirosław-grabowska and niska, 2007b; kołaczek et al., 2012). therefore, only the palynological analysis document the eemian and early vistulian age of the solniki paleolake deposits. the present work refers to the stratigraphic nomenclature of poland, where the vistulian, eemian and wartanian periods correspond to marine isotope stages (mis) 2-5d, 5e and 6, respectively (pisias et al., 1984). the vistulian is subdivided into early (mis 5d-a), plenivistulian (lower mis 4, middle mis 3 and upper plenivistulian lower part of mis 2), and late vistulian (upper part of mis 2 and lowermost part of mis 2). we correlated eemian interglacial with ipswichian, mikulinian and sangamonian mis-5e, the vistulian with the western europe weichselian glacial period (syn. würmian, valdaian, devensian, wisconsin mis-5e–mis-1), and the wartanian with the late saalian glacial period (syn. riss ii, warthe glaciation, wolstonian) (gibbard and kolfschoten, 2004; pisias et al., 1984). the eemian interglacial is correlated with the marine isotope stage – mis 5e (oxygen isotope stages – ois 5e) and the early vistulian is correlated with mis 5d-c (mangerud, 1989; shackleton et al., 2003). cladocera analysis cladocera analysis was performed for the sediment section 3.00-11.00 m depth, with a resolution of 5-10 cm (overall 99 samples). the samples for cladocera analysis were taken from the same sections analysed for pollen, but the resolution was constrained by the scarcity of the material left by pollen analyses. the samples (1 g) were prepared according to standard procedures (frey, 1986). after removing the carbonates using 10% hcl, the samples were heated at 60°c with 10% koh for 20 min. the residue was washed with distilled water and sieved through a 40-μm-mesh sieve. the samples were coloured with safranine and identified with a light microscope at 100×, 200× and 400× magnification. when possible, a minimum of 200 remains of cladocera were analysed, using five to seven slides from each sample. all cladocera remains were counted in each slide (head shield, shell, postabdomen, postabdominal claws, antennules and others), and the conversion of counted remains to the total fig. 1. location of the solniki paleolake. no n c om me rci al us e o nly eemian/vistulian development of a polish paleolake 209 number of individuals was based on the most abundant type of remains for each taxon according to frey (1986). the total number of cladocera specimens in 1 g of sediment and the species diversity were assessed according to the shannon-wiener diversity index and plotted using the polpal software (nalepka and walanus, 2003; fig. 2). identification of cladocera remains and ecological interpretation were carried out according to frey (1958, 1962), goulden (1964), szeroczyńska (1985, 1989), hofmann (1986, 2000), korhola (1990), duigan (1992), flössner (2000), szeroczyńska and sarmaja-korjonen (2007), rybak and błędzki (2010), and nevalainen et al. (2013), and nevalainen and luoto (2016). all cladocera species were classified into planktonic (pelagic) and littoral (meiobenthic) groups, with the latest group sub-categorised into three assemblages with different habitat preference, i.e. bottom-dwellers, species associated with vegetation and species restricted to vegetation (flössner, 1964; whiteside, 1970; hofmann, 1987; whiteside and swindoll, 1988, korhola, 1990). the species chydorus sphaericus (o.f. müller) and bosmina longirostris (o.f. müller) were assign to booth groups: littoral and open water zones because of their ability to migrate (whiteside, 1970). results the results of qualitative and quantitative cladocera analyses of the solniki sediments are presented as absolute number of individuals in fig. 3. the subfossil cladoceran fauna of the solniki paleolake are represented by 17 species belonging to the families bosminidae, chydoridae, sididae and daphniidae (fig. 3). fourteen of the identified species typically dwell within the littoral zone (chydoridae). three species occur mainly in the pelagic zone (bosminidae, daphniidae). most of the remains belong to the families bosminidae, with dominance of eubosmina coregoni (baird), and chydoridae, with a predominance of alona affinis (leydig) and camptocercus rectirostris (schoedler). these species prefer littoral zone in clear and calm waters in relation to the presence of aquatic plants (flössner, 1972). resting eggs of the ceriodaphnia sp. have also been identified in the deposits. tab. 2. correlation of local pollen assemblage zones (l paz) from the solniki profile with regional pollen assemblage zones distinguished for poland (r paz) and chronostratigraphical units and oxygen isotope stages (modified from mirosław-grabowska et al., 2015). local pollen assemblage regional pollen assemblage chronostratygraphy oxygen isotope zonesl paz according zonesr paz according stages according to kupryjanowicz, 2008 to mamakowa, 1989 to pisias et al., 1984 (ois) s-12 artemisia-poaceae ev3 graminae-artemisia ev3 rederstall stadial 5b betula nana s-11 pinus-betula ev2 betula-pinus brørup s.s. ev2 brørup s-10 artemisia-betula cool osc. interstadial 5c s-9 betula amersfoort sensu lato s-8 artemisia-cyperaceae ev1 graminae-artemisia ev1 herning stadial 5d poaceaebetula nana betula nana s-7 pinus e-7 pinus late 5e s-6 picea-alnus-carpinus-pinus e-6 picea-abies-alnus s-5 carpinus-alnus e-5 carpinus-corylus-alnus middle 5e s-4 corylus-alnus-tilia e-4 corylus-quercus-tilia s-3 quercus-ulmus-fraxinus e-3 quercus-fraxinus-ulmus s-2 pinus-salix-quercus-ulmus e-2 pinus-betula-ulmus early 5e s-1 pinus-betula-picea e-1 pinus-betula tab.1. lithology of the solniki profile (modified from kupryjanowicz, 2008). depth (m) sediment description 3.00-3.90 peaty silt with clay, dark grey; homogeneous, with plant detritus 3.90-5.60 organic silt with sand, slightly clayey, black 5.60-8.20 organic clayey silt, brown-black 8.20-10.00 organic clayey silt with small admixture of caco3 brown 10.00-10.60 organic silt, black 10.60-11.00 peat, slightly clayey with traces of sand, black e ar ly v is tu lia n e em ia n in te rg la ci al no n c om me rci al us e o nly 210 m.m niska different frequency and concentration enables to distinguish six cladocera assemblage zones (caz) in the solniki profile (figs. 2 and 3). zone caz i (11.00-9.90 m) up to 12 species per samples are present in this section, and total number of cladocera individuals is around 950 ind g–1 the predominant species are eubosmina coregoni (max. 76% of total cladocera abundance, figs. 2 and 3), a. affinis (max. about 32%), and eurycercus lamellatus (o.f. müller) (max. about 25%). ch. sphaericus and sida crystallina (o.f. müller) are also present. pelagic species (e. coregoni and ceriodaphnia sp. toward the end of this zone) and taxa related to aquatic vegetation are dominant. finally, two species of the group connected with the bottom are present in this zone, i.e. monospilus dispar (sars) and pleuroxus uncinatus (baird) (vegetation/bottom species). the central part of this zone shows the highest diversity index (0.78; fig. 2). zone caz ii (9.90-8.15 m) due to differences in the frequency of species and specimens zone ii, is divided into two subzones. subzone caz ii a (9.90-9.35 m) the number of individuals increases up to 2100 ind. g– fig. 2. diagram of the total number of cladocera specimens and species richness diversity and ratio of planktonic to littoral form p/l in the sediments of the of solniki profile. l paz, local pollen assemblage zones (according to kupryjanowicz, 2008); caz, cladocera assemblage zones; exl., exclude; b. long, bosmina longirostris; ch. sph., chydorus sphaericus; o, oligotrophy; m, mesotrophy; e, eutrophy; the arrows indicate increase and decrease in the lake water level. explanations of lithology in fig. 3. no n c om me rci al us e o nly eemian/vistulian development of a polish paleolake 211 1 of sediments (fig. 2), while species richness decreases down to 8. as in the previous zone, three species are dominant: e. coregoni (max. about 57%), a. affinis (27%) and c. rectirostris (max 19%, fig. 3). an increase in the share of species from the group restricted to vegetation is recorded in this zone, while only e. coregoni represents the open water group. the diversity index is low (0.69-1.1). subzone caz ii b (9.35-8.15 m) the total number of individuals decrease down to 400 ind. g–1, while species richness increases up to 11 (fig. 2). a. affinis and e. coregoni are the dominant species. species of the group associated with water plants are the most abundant. the diversity index is the same as in the previous subzone 0.68 (max 1.1). zone caz iii (8.15-6.25 m) zone iii is distinguished due to the similarity of species composition in the samples and distinct separation from the rest of the zones. however, due to differences in the number of individuals of each species is divided into 6 subzone. subzone caz iii a (8.15-7.68 m) species richness increases (up to 17), as well as the abundance of species of all groups (fig. 3). the total number of individuals increases to a max. of about 1400 ind g–1 (fig. 2). bosminidae are dominant, with e. coregoni (max about 50%), b. longirostris, and species associated with vegetation, such as a. affinis, e. lamellatus, small alona and s. crystallina. species connected with the sediment, such as alona quadrangularis (o.f. müller) and leydigia leydigi (schoedler) are found for the first time in the sediment record, as well as graptoleberis testudinaria (fischer), small alona and alonella exigua (lilljeborg), which are related to connected high nutrient levels. fig. 3. diagram of the absolute number of cladocera specimens in the sediments of solniki profile. caz, cladocera assemblage zones. no n c om me rci al us e o nly 212 m.m niska the diversity index increases up to 0.74. subzone caz iii b (7.68-7.36 m) total abundance of individuals father increases (up to 5400 ind. g–1), while the species richness slight decrease (15; fig. 2). bosminidae family predominates also in this zone with e. coregoni and b. longirostris (up to 2000 ind g–1). the abundance of littoral species a. affinis, s. crystallina, small alona, e. lamellatus and ch. sphaericus increases. the increase in species like m. dispar and a. exigua indicates mesotrophic water status, while l. leydigi and acroperus harpae (baird) disappear. the presence of both cold-tolerant (a. affinis, e. lamellatus, ch. sphaericus) and warm requiring species (c. rectirostris, p. uncinatus) are present, and diversity index increases up to 0.95. subzone caz iii c (7.36-7.14 m) species richness decreases down to 11 in this zone, as also the total number of cladocera (maximum = 1200 ind g–1). b. longirostris (about 50%, fig. 3), e. coregoni and a. affinis predominate, while some species tolerating cold water are also present. the diversity index decreases again down to 0.84. subzone caz iii d (7.14-6.88 m) the conditions for zooplankton development improve as indicated by the increase in the abundance of both littoral and planktonic species (similarly as in the caz ii b subzone). values of diversity index are above 1.1, and both species richness (up to 15) and total claodcera abundance (up to 4950 ind g–1) increase. dominant species are e. coregoni (max. 2000 ind g–1), b. longirostris and a. affinis. species preferring high water nutrient (p. uncinatus, l. leydigi, g. testudinaria) and warmer water (c. rectirostris and p. uncinatus) reach higher frequencies. nevertheless both cold-tolerant and warm water-requiring species are present. subzone caz iii e (6.88-6.72 m) this subzone is characterized by a decreasing abundance of specimens (to 1150 ind g–1; fig. 2) and species richness (12). species associated with water plants (a. affinis, a. harpae, e. lamellatus and ch. sphaericus) are dominant, and b. longirostris and e. coregoni are also numerous. the diversity index decreases down to 0.94. subzone caz iii f (6.72-6.25 m) both species richness and total cladocera abundance decrease (down to 13 and 1350 ind g–1, respectively; fig. 2). littoral species such as a. affinis, a. harpae, a. quadrangularis, and ch. sphaericus dominate in this zone. e. coregoni is the dominant pelagic species (fig. 3). except for a. quadrangularis, there are no species associated with the bottom. taxa preferring a high temperature of water (c. rectirostris, p. uncinatus, g. testudinaria) are identify in the second part of this subzone. zone caz iv (6.25-5.71 m) species richness and total abundance of individuals further decrease down to 9 and 330 ind g–1 (fig. 2). only cold-tolerant species (i.e., a. affinis, ch. sphaericus, e. coregoni and b. longirostris) remain in the lake. the disappearance of species preferring warm waters (pleuroxus sp., c. rectirostris) disappear and the diversity index varies from 0.00 to 0.88. zone caz v (5.71-3.72 m) this zone is distinguished due to the similarity of species composition in the samples and distinct separation from the rest of the zones. however, due to differences in the number of individuals of each species is divided into four subzone. almost all species identified in the previous zones are present again in this zone. subzone caz v a (5.71-5.21 m) this zone is characterized by the increase in frequency of both littoral and planktonic species. species richness increases up to 16 (fig. 2) and the total number of individuals up to 2300 ind g–1. dominant species are e. coregoni and b. longirostris, a. affinis, small alona, and a. quadrangularis. both thermophilic (e.g., p. uncinatus, and c. rectirostris) and cold water-tolerant species (e.g., a. harpae, e. lamellatus), as well as species requiring more fertile water (e.g., l. leydigi, b. longirostris) are present. diversity index increases up to 1.05. subzone caz v b (5.21-5.02 m) total cladocera abundance drops again down to 970 ind g–1. species richness decreases down to 13. cold-tolerant species (ch. sphaericus, a. affinis, small alona) are dominant (fig. 3), and diversity index is low (max. 0.77). subzone caz v c (5.02-4.18 m) both species richness and total number of individuals increased (up to 17 species and 2050 ind g–1; fig. 2). the most abundant species are ch. sphaericus, e. coregoni, c. rectirostris and a. affinis. the diversity index rises up to 1.0. subzone caz v d (4.18-3.72 m) the abundance of all species significantly increases and the total number of cladocera reaches 4300 ind g–1. on the contrary, species richness decreases down to 15. species associated with water plants (a. affinis, c. rectirostris) and with the bottom (a. quadrangularis and ch. no n c om me rci al us e o nly eemian/vistulian development of a polish paleolake 213 sphaericus) are dominant, but also species requiring either warm waters (c. rectirostris, g. testudinaria and p. uncinatus) or high nutrient level (b. longirostris, a. quadrangularis, l. leydigi and ch. sphaericus) are present. s. crystallina and alonella nana (baird) disappear in this zone. the diversity index remains around 1.0. zone caz vi (3.72-3.00 m) the abundance of all species decreases, except for ch. sphaericus and b. longirostris, which are predominant. total individual number drop down to 1300 ind g–1, and species richness decreas down to 6. the presence of coldwater tolerant species, such as a. harpae, ch. sphaericus, small alona and b. longirostris, are present, and the diversity index drops down to 0.54. discussion the subfossil cladocera analysis conduct on the solniki profile provide information on the paleolake evolution during the eemian interglacial and the early vistulian glaciation (ev1-ev3). the remains are well preserve and abundant considering the old age of the sediments, and the cladocera species composition is comparable to what found in different central european lakes (szeroczyńska and zawisza, 2007) and eemian paleolakes, such as kaliska, studzieniec, ruszkówek, kuców iic (niska, 2012), sławoszewek (pawłowski, 2011), imbramowice (mirosław-grabowska and gąsiorowski, 2010), and in herning paleolake (frey, 1962). lake evolution of the solniki paleolake during the eemian-vistulian solniki paleolake develop in a melt depression at the end of late saalian glaciation, while the lake sedimentation starts at the onset of the early eemian (kupryjanowicz, 2008), when the improvement of climatic conditions and the weakening of periglacial processes chang the vegetation surrounding the lake and allow the development of the biological community within the new lake. early eemian local pollen assemblage zone (l paz) s-1 (pinusbetula-picea), regional pollen assemblages zone (r paz) e1, l paz s-2 (pinus-salix-quercus-ulmus), r paz e2, begging of caz i. during the early eemian, the area surrounding the lake initially develop boreal pine forests with quite a large share of spruce (l paz s-1). this may indicate a cool, relatively humid climate. in the period corresponding with zone l paz s-2, in the place of spruce enter more demanding trees in terms of temperature, such as oak and elm. the willow, inhabiting moist land (kupryjanowicz et al., 2004, kupryjanowicz, 2008) also reach its maximum growth. no cladocera remains are find in the studied sediment profile at the very beginning of the early eemian, due to the peaty condition of the early stage of the new lake, while the following samples are characterize by the presence of ch. sphaericus and a. affinis (caz i), which are consider as pioneer species due to their tolerance to low water temperature and wide tolerance to environmental conditions (whiteside, 1970; hofmann, 2000). similar species composition at the beginning of the early eemian is observed also in other polish lakes, such as starowlany paleolake and kubłowo paleolake (niska and roman, 2014; niska and kołodziej, 2015), as well as in central and northern european lakes (szeroczyńska, 1985; sarmaja-korjonen and alhonen, 1999; gąsiorowski and szeroczyńska, 2004). middle eemian l paz s-3 (quercus-ulmus-fraxinus) r paz e3, l paz s-4 (corylus-alnus-tilia) r paz e4, l paz s-5 (carpinus-alnus) r paz e5, caz i, caz ii, caz iiia. the interglacial climatic optimum starts with oak-elm forests with a fairly large share of ash (fraxinus) (l paz s-3, e3). they are most likely riparian forests, where the role of the pine is significantly lower and hazel (corylus) gradually spread in the undergrowth. multispecies of thermophilic deciduous forests successively develop (l paz s-4, e4), with hazel (corylus) and lime tree (tilia) reaching their optimum growth, and elm (ulmus), oak (quercus) and hornbeam (carpinus), and rarely (fraxinus) as accompanying species. alder forests are growing on the wetlands areas. all types of forests show low share of non-arboreal pollen grains (nap), which did not exceed 2%. during the middle eemian the peat depositions are replace by silt. moreover the cladocera composition show a change in specie assemblages. in fact, despite the constant dominance of the pioneer specie a. affinis at the beginning of the zone (caz iia), several cladocera species are also identify in this part of the profile. a. nana and e. coregoni appear for the first time suggesting oligo/mesotrophic conditions. the presence of the planktonic e. coregoni indicates an increase in water depth (hofmann, 1993). moreover, the identification of p. uncinatus, c. rectirostris and m. dispar reveal a progressive increase in water temperatures (poulsen, 1944; frey, 1958), while the presence of species associated to water plants (i.e., a. harpae, a. affinis, a. nana, e. lamellatus) suggest a development of the macrophyte zone. in the next period (l paz s-5 r paz e5), forests dominated by hornbeam absolutely prevail in the majority of habitats these hornbeam forest has also initially a relatively high share of hazel, lime tree and elm. spruce (picea) appears and gradually spread, the role of pine and birch also no n c om me rci al us e o nly 214 m.m niska increas the humid places are dominated by alder forests. the recorded pollen trees suggest hotter and longer summers than in the zone l paz s-4, and still mild winter (kupryjanowicz, 2008; niska, 2008). mainly cold water tolerant species, i.e. a. harpae, e. lamellatus and ch. sphaericus are identify in this sediment section. however, the increase in c. rectirostris, p. uncinatus and m. dispar suggests an improvement in climatic conditions the sedimentation of silt is replace by gyttja, which may indicate a deepening of the reservoir, in agreement with the increase in abundance of pelagic species, such as e. coregoni. the next stage (caz ii b) bring a deterioration of habitat conditions in the reservoir. initially only the summer temperature decreased (kupryjanowicz, 2008), but the palynological analysis indicates a later winter tightening and a climate drying (kupryjanowicz et al., 2004). at that time, a significant reduction in the abundance of cladocera is observed in the other eemian sites in poland, e.g., kaliska, ruszkówek and kubłowo (mirosław-grabowska and niska, 2007a; mirosławgrabowska et al., 2009; niska and roman, 2014). in paleolake herning, a drop in the abundance of cladocera is noted at the end of r paz e4 and the beginning of r paz e5 (frey, 1962). the observed decrease in total cladocera abundance in this sediment section, together with the dominance of a. affinis and e. coregoni, confirm the decrease in water temperature, and suggest a correspondent decrease in water level and in the lake trophic status. the degradation of the living conditions in the lake could be associate not only to the low water temperatures but also to the reduction in the water level. in fact kupryjanowicz (2007) show that most of the eemian paleolakes from the northern podlasie region dried at that time, producing a terrestrialization of several shallow lakes and bogs. at that time, a significant reduction in the cladocera abundance is observed in the other eemian sites in poland, e.g. kaliska, ruszkówek and kubłowo (mirosław-grabowska and niska, 2007a; mirosław-grabowska et al., 2009; niska and roman, 2014). at the end of this period, a further cooling occurs with a simultaneous increase in climate humidity (kupryjanowicz et al., 2004). the slight increase in cladocera remains (caz iiia) and in particular in e. coregoni suggested a new stage of high water depth, while the presence of species associated with water plants, such as a. affinis, e. lamellatus, and s. crystallina, indicat an expansion and a development of the macrophytes zone. moreover, the identification of m. dispar and a. exigua in this zone suggest mesotrophic conditions (adamska and mikulski, 1968). late eemian l paz s-6 (picea alnus-fraxinus-pinus) r paz e6, l paz s-7 (pinus) r paz e 7, caz iii b-f. the late eemian interglacial period is characterized by the maximum development of spruce. pine, fir (abies) and larch (larix) are dominant, along with some deciduous trees such as birch (betula) and hornbeam (carpinus). these trees can form different types of coniferous and mixed forests. also, alder (alnus) is present in pretty large quantities (kupryjanowicz et al., 2004; kupryjanowicz, 2008; niska 2008). the total cladocera abundance and species diversity indicates that it was one of the most favourable periods for the development of zooplankton in the lake. during the late eemian the cladocera community register rapid increases (caz iii b-d-f) alternate to drastic decreases in species diversity and abundance (caz iii c-e), suggesting a continuous change from favourable to unfavourable lake conditions for the zooplankton development. in particular, in caz iii b-d-f, the simultaneous presence of the planktonic species e. coregoni and b. longirostris indicate an increase in the trophic level (boucherle and zullig, 1983; szeroczyńska, 1985, 1991; frey, 1986; nauwerck, 1991; dimante-deimantovica et al., 2012), which is also confirmed by the presence of m. dispar and a. exigua. moreover the identification of c. rectirostris, p. uncinatus and g. testudinaria reveals an increase in water temperatures (szeroczyńska, 1985; hofmann, 2000), while remains of a. harpae, a. quadrangularis and l. leydigi suggest a continuous development of the macrophyte zone. the caz iii c-e periods are characterized by low cladocera abundances and by the presence of the cold water tolerant species a. affinis, e. lamellatus and a. nana, suggesting deteriorated ecological conditions due to the low water temperatures. similar cladocera trends during the late eemian is documented in the other paleolakes, e.g., kuców iic (niska, 2008), kaliska and kubłowo from central poland (mirosław-grabowska and niska, 2007a; niska and roman, 2014), indicating a comparable regional influence on the cladocera community by regional climate development. on the other hand, although the cladocera species diversity and abundance increase in the herning paleolake as in the other paleolakes, changes in species frequency are not so significant due to the smaller differences in the climate condition and more moist (oceanic) climate (mirosław-grabowska et al., 2016). the pollen assemblages in solniki paleolake confirm that the late eemian is a period characterized by short-term colder and warmer events (kupryjanowicz, 2008). early vistulian, herning stadial l paz s 8 (artemisia-cyperaceae-poaceaebetula nana) – r paz ev1, caz iv the herning stadial was the first post-eemian cold period in northern podlasie. the solniki core are characterize by a high content of clay and sand, which indicate a diluvial and/or eolic genesis of these deposits in relation no n c om me rci al us e o nly eemian/vistulian development of a polish paleolake 215 to a stage of climate cooling and increasing humidity. at that time, vegetation of the studied region is predominantly represented by tundra, composed by willow clusters, possibly enriched with shrub birches. trees were likely present only in form of small patches of pine and birch(kupryjanowicz, 2008). in this zone, a pronounced decrease of the subfossil cladoceran fauna are observed. as registered also in other polish lakes between the end of the eemian interglacial and the beginning of the vistulian (i.e., kubłowo, ustków, rzecino; niska, 2015). cladocera assemblages is composed by the sole species tolerating low water temperatures, such as a. affinis, ch. sphaericus, b. longirostris and e. coregoni, suggesting a significant reduction of the water temperatures. early vistulian, brørup l paz s-9 (betula), amersfoort, s-10 (artemisia-betula) (cool oscillation), s-11 (pinus-betula), brørop sensu stricto, r paz ev2, caz v a-d in the older part of the brørup sensu lato interstadial (s-9 betula l paz), which may be correlated to the amersfoort interstadial distinguished in western europe, signs of a climate improvement are observed (kupryjanowicz, 2008). birch forests spread at the cost of areas occupied by herbaceous plant communities. dry habitats remain covered by steppe communities, whereas wet habitats are still occupied by grass communities similar to contemporary damp meadows. pine trees (pinus sylvestris type (linneaus) become an essential component of birch forests or even formed separate assemblages. at the beginning of the brørup (early vistulian, caz v a) cladocera abundances and diversity increase. the presence of thermophilous species such as p. uncinatus and c. rectirostris indicates higher water temperatures, although species tolerating cool water (e.g., a. harpae and e. lamellatus) are still present. the increase in l. leydigi and b. longirostris suggest an increase in the lake nutrient levels (alhonen, 1970; szeroczyńska, 2002). the high abundance of the family chydoridae associated to aquatic vegetation may indicate a new stage of macrophyte development. finally, the increase in e. coregoni may reveal another rise in water level. as outlined by the pollen analysis, the climate in this period is more humid and characterize by higher rainfall events (kupryjanowicz, 2008), which may confirm the hypothesis of a higher lake water depth. the period of favourable conditions for the cladocera development is interrupted by a rapid cooling stage (l paz s-10, caz v b), which, in the northern podlasie, is reflected by a spread of the area covered with cold steppe communities lush with artemisia, chenopodiaceae, anthemis type, caryophyllaceae undiff., and some grasses (kupryjanowicz, 2008). the subfossil cladoceran community are represent in this stage only by species tolerant of cold water, such as a. quadrangularis, a. affinis, ch. sphaericus, b. longirostris and e. coregoni. despite a short period of cold water temperatures characterized by a decrease in cladocerea abundances (caz v b), the favourable conditions for the cladocera growth continue throughout the entire brørup interstadial (caz v c-d). this period are characterize by species preferring warm water temperatures and the replacement of e. coregoni by b. longirostris suggest a reduction in the lake water level and an increase in the lake trophy (hofman 1987; szeroczyńska, 1998b; nevalainen et al., 2013). the high cadocera abundance and diversity in this period reveal an improvement of the water living conditions, which is also observed in kubłowo lake (niska and roman, 2014). the pollen analysis by kupryjanowicz (2008) highlight a boreal forest spread, which confirmed the improving of the climate conditions. early vistulian, rederstall stadial l paz s-12 (artemisia-poaceae) r paz ev3, caz vi the cold continental climate of the rederstall stadial led to the spreading of cold steppe communities. betula nana type (linneaus) and shrub salix form a dwarf shrub tundra, while forest communities declined. in the rederstall stadial (early vistulian) the cladocera abundance and diversity decrease (caz vi). cold water tolerant species are present, e.g., a. harpae, and ch. sphaericus, as well as species indicating high concentration of organic matter in the water column, e.g., b. longirostris, small alona, a. quadrangularis, and ch. sphaericus (duigan, 1992). the dominance of ch. sphaericus and b. longirostris suggests eutrophic lake conditions and general deterioration of the ecological quality (szeroczyńska 1991). in fact, both taxa are widely tolerant to environmental stress, and they are often the last cladoceran inhabitants of shallowing water bodies in palaeorecords (nováková, 2005). cladocera record for the rederstal stadial period are available in poland only for other two lakes: kubłowo (miroslaw-grabowska at al., 2016) and ustków (kołaczek et al., 2016). in the kubłowo profile wide tolerant species ch. sphaericus, a. quadrangularis and a. nana are dominant, as similarly recorded also in solniki paleolake. on the other side, cladocera community in ustków profile appears to be dominated by b. longirostris, ch. sphaericus and small alona, which are species tolerating high environmental stress (duigan, 1992). the cladocerea assemblages of this two paleolake include also species preferring high water temperature (i.e., pleuroxus sp., m. dispar, c. rectirostris), while the lack of thermophilic species in the solniki paleolake could be the result of a more severe continental climate occurring in this area. no n c om me rci al us e o nly 216 m.m niska water level changes the planktonic:littoral (p:l) ratio of cladocera was used for the first time by alhonen (1970) to investigate lake level changes during the holocene. however, the interpretation of p/l ratios should take into account that many other environmental variables, in addition to lake depth, can affect the share of pelagic and littoral of subfossil cladoceran assemblages (gąsiorowski and hercman, 2005;nevalainen et al., 2008). the large average share of deep-water species (48%) in the solniki profile suggests a persistent presence of a well-developed pelagic zone in the paleolake. nevertheless, the depth profile of the p:l ratio shows a general predominance of shallow-water groups, such as chydoridae and sididea, which suggests the constant presence of a shallow littoral zone. during the middle and late eemian interglacal period the p:l ratio reveals a rise in planktonic species, such as e. coregoni, b. lonigrostris and ceriodaphnia spp, suggesting an increase in lake depth. the drastic decrease in littoral species and the presence of the sole planktonic taxa e. coregoni during the herning stadial suggest that solniki lake is still a deep lake during this stage. however, this result has to be considered as a fake since the l:p ratio did not take into account the general decrease in total cladcocera abundance and species richness. this provides a confirmation of the necessity for a carefully interpretation of the p:l cladocera ratio for paleoecological reconstructions. on the other hand, the early vistulian period shows a prevalence of littoral species, which can be read as a gradual lowering of the water level in the paleolake, which may have contributed to enlarge the shallow littoral zone. lake trophic evolution cladocera assemblages indicate that solniki paleolake was nutrient poor for most of its existence. in fact, during the middle eemian the paleolake is dominated by species preferring oligotrophic conditions. an increase in the lake trophic condition, likely up to a mesoc-eutrophic level, occur during the late eemian, as indicated by the presence of m. dispar, a. exigua and b. longirostris (adamska and mikulski, 1968). the decrease in cladocera abundances during the end of the late eemian interglacial and the herning stadial suggests a new long stage of low nutrient concentrations. on the other hand, the presence of ch. sphaericus and of b. longirostris in the final stage of the lake existence, reveal a second increase in nutrient concentrations till eutrophic conditions during the brørup interstadial and the rederstall stadial. conclusions the unique character of the paleolake at solniki consists in its very long, uninterrupted sedimentation that occurred during the entire eemian interglacial and early vistulian. the detail cladocera analysis showed that: the cladocera remains preserve in sediments of the solniki paleolake, despite their age, are well preserve. their species composition are not substantially differ from those found in of other eemian paleolakes and in contemporary central european lakes. the species richness is slightly lower than in contemporary lakes but very similar to that of eemian paleolake starowlany (ne poland). the most favourable periods for cladocera development are identify at the beginning of the eemian interglacial during some stages of the late eemian and during the brørup interstadial , probably in relation to higher water level, warmer temperatures and moderate nutrient levels. such favourable stages were also recorded in the other studied eemian lakes from central europe. unfavourable conditions for cladocera community occur during the beginning of the eemian interglacial, the middle eemian (caz ii b) and the herning stadial (caz iv), probably due to low lake water temperatures and nutrient level. several major fluctuations in the lake water level are observe during the entire studied period, in relation to the alternation of warm and wet and dry and cold climate conditions. cladocera assemblages suggest oligotrophic to oligomesotrophic conditions at the beginning of the lake existence, which are follow by a steady increase in nutrient concentrations till the final lake stage. major changes in subfossil cladocera frequency track the alternation of warmer and cooler temperatures throughout the entire studied period. warmer water temperatures are recorded during the early eemian, the beginning of the middle eemian and the brørup interstadial, while colder temperatures are identified in the late eemian, in herning stadial and again in the rederstal stadial. acknowledgments i would like to express my thanks to prof. mirosława kupryjanowicz for the very interesting discussions. the cladocera analysis was supported by the foundation for polish science, bridge program 2012 – project: “reconstruction of the development of lake environmental in the eemian interglacial based on subfossil cladocera (crustacea) analysis”. fieldworks, pollen and part of the cladocera analysis were supported by the ministry of science and higher education, project 2po4e 02329. references no n c om me rci al us e o nly eemian/vistulian development of a polish paleolake 217 adamska a, mikulski js, 1969. cladocera remains in the superficial sediments of lakes as a typologic indicator. z. naukowe umk 25, prace stacji limnolog w iławie 5:41-48. alhonen p, 1970. on the significance of the planctonic/litoral ratio in the cladoceran stratigraphy of lake sediments. comentat. biol. 57:3-34. börner a, hrynowiecka a, kuznetsov v, stachowicz-rybka r, maksimov f, grigoriev v, niska m, moskal-del hoyo m, 2015. palaeoecological investigations and 230th/u dating of eemian interglacial peat sequence of banzin (mecklenburg-western pomerania, ne-germany). quatern. int. 386:122-136. boucherle mm, züllig h, 1983. cladoceran remains as evidence of change in trophic state in three swiss lakes. hydrobiologia 103:141-46. brodersen kp, whiteside mc, lindegaard c, 1998. reconstruction of trophic state in danish lakes using subfossil chydorid (cladocera) assemblages. can. j. fish. aquat. sci. 55:1093-1103. bruj m, krupiński km, 2001. [biogeniczne osady jeziorne interglacjału eemskiego w żeliszczewie na wysoczyźnie siedleckiej].[article in polish]. przegląd geologiczny 49:583-543. cheddadi r, mamakowa k, guiot j, de beaulieu jl, reille m, andrieu v, granoszewski w, peyron o, 1998. was the climate of the eemian stable? a quantitative climate reconstruction from seven european pollen records. palaeogeogr. palaeoclim. palaeoecol. 143:73-85. cohen as, 2003. paleolimnology: the history and evolution of lake systems. oxford university press, new york. dimante-deimantovica i, skute a, skute r, 2012. vertical variability of pelagic zooplankton fauna in deep latvian lakes, with some notes on changes in ecological conditions. eston. j. ecol. 61:247264. dickinson ka, 1959. the ostracoda and cladocera of the humboldt formation in northeastern nevada. ms thesis (unpublished), university of minnesota. dobracka e, winter h, 2001. [stanowisko osadów interglacjału eemskiego w profilu otworu rzycino (wysoczyzna łobezka)].[article in polish]. proceedings 8th conf. “stratygrafia plejstocenu polski” jarnołtówek. duigan ca, 1992. the ecology and distribution of the littoral freshwater chydoridae (branchiopoda, anomopoda) of ireland with taxonomic comments on some species. hydrobiologia 241:1-70. duigan ca, birks hh, 2000. the late-glacial and early-holocene palaeoecology of cladoceran microfossil assemblages at kråkenes, western norway, with quantitative reconstruction of temperature changes. j. paleolim. 23:67-76. erni a, forcart l, harri h, 1943. [fundstellen pleistocaener fossilien in der “hochterrasse” von zeli (kt. luzem) und in der morane der größten eiszeit von auswil bei rohrbach (kt. bern)].[article in german]. eciogae geol. hel. flössner d, 1964. [zur cladoceranfauna des stechlin-gebietes. ii. őkologische untersuchungen über die litoralen arten].[article in german]. limnologica 2:35-103. flössner d, 1972. branchipoda, branchiura. tierwelt deutschl. 60:1-501. flössner d, 2000. [die haplopoda und cladocera (ohne bosminidae) mitteleuropas].[book in german]. backhuys publishers, leiden: 428 pp. frey dg, 1958. the late glacial cladoceran fauna of a small lake. arch. hydrobiol. 54:209-275. frey dg, 1962. cladocera from the eemian interglacial of denmark. j. paleontol. 36:1133-1154. frey dg, 1986. cladocera analysis, p. 667-692. in: b.e. berglund (ed.), handbook of holocene palaeoecology and palaeohydrology. j. wiley & sons, chichester. gąsiorowski m, 2002. changes in cladocera (crustacea) assemblage of lake kruklin (masurian lake district) caused by an artificial drop in water level. limnol. rev. 2:131-136. gąsiorowski m, szeroczyńska k, 2004. abrupt changes in species structure of genus bosmina during the history of the ostrowite lake (northern poland). hydrobiologia 526:137-144. gąsiorowski m, hercman h, 2005. recent sedimentation and eutrophication of kruklin lake after artificial drop in waterlevel in the middle of 19th century. stud. quatern. 22:17-25. gibbard p, kolfschoten t, 2004. the pleistocene and holocene epochs, p. 441-452. in: f.m. gradstein, j.g. ogg and a.g. smith (eds.), a geologic time scale. cambridge university press, cambridge. goulden ce, 1964. the history of the cladoceran fauna of esthwaite water (england) and its limnological significance. arch. hydrobiol. 60:1-53. granoszewski w, 2003. late pleistocene vegetation history and climatic changes at horoszki duże, eastern poland: a palaeobotanical study. acta palaeobotanica suppl. 4:3-95. harmsworth rv, 1968. the developmental history of blelham tarn (england) as shown by animal microfossils, with special reference to the cladocera. ecol. monogr. 38:223-241. hann bj, karrow pf, 1984. pleistocene paleoecology of the don and scarborough formations, toronto, canada, based on cladocera microfossils at the don valley brickyard. boreas 13:377-391. hann bj, karrow pf, 1993. comparative analysis of cladoceran microfossils in the don and scarborough formations, toronto, canada. j. paleolimnol. 9:223-241. hofmann w, 1986. developmental history of the grosser plöner see and the schöhsee (north germany): cladoceran analysis, with special reference to eutrophication. arch hydrobiology. suppl. 74:259-287. hofmann w, 1987. stratigraphy of cladocera (crustacea) and chironomidae (insecta: diptera) in three sediment cores from the central baltic sea as related to paleo-salinity. int. rev. ges. hydrobiol. 72:97-106. hofmann w, 1993. late-glacial/holocene changes of the climatic and trophic conditions in three eifel maar lakes, as indicated by faunal remains. i. cladocera. lect. notes earth sci. 49:393-420. hofmann w, 2000. response of the chydorid faunas to rapid climatic changes in four alpine lakes at different altitudes. palaeogeogr. palaeoclimatol. palaeoecol. 159:281-292. jeppesen e, jensen jp, skovgaard h, hvidt chb, 2001. changes in the abundance of planktivorous fish in lake skanderborg during the past two centuries a palaeoecological approch. palaeogeogr. palaeoclimatol. palaeoecol. 172:143-152. jeppesen e, kronvang b, olesen je, søndergaard m, hoffmann cc, andersen he, lauridsen tl, liboriussen l, meerhoff m, beklioglu m, özen a, 2011. climate change effect on nitrogen loading from catchment in europe: implications for no n c om me rci al us e o nly 218 m.m niska nitrogen retention and ecological state of lakes and adaptations. hydrobiologia 663:1-21. jessen k, milthers v, 1928. stratigraphical and paleontological studies of interglacial fresh-water deposits in jutland and north-west germany. danm. geol. unders. 48 1-380. kadota s, 1975. a quantative study of the microfossils in a 200meter-long core sample from lake biwa. paleolimnol. lake biwa jpn pleistocene 85:354-367. korhola a, 1990. paleolimnology and hydroseral development of the kotasuo bog, southern finland, with special reference to the cladocera. ann. acad. sci. fenn. 155:5-40. korhola a, rautio m, 2001. cladocera and other branchiopod crustaceans, p. 5-41. in: j.p. smol, h.j.b. birks and w.m. last (eds), tracking environmental change using lake sediments. 4. zoological indicators. kluwer academic publishers, dordrecht. kołaczek p, karpińska-kołaczek m, petera-zganiacz j, 2012. vegetation patterns underclimate changes of the eemian and early weichselian in central europe inferred from a palynological sequence from ustków (central poland). quatern. int. 268:9-20. kołaczek p, niska m, mirosław-grabowska j, gałka m, 2016. periodic lake-peatland shifts under the eemian and early weichselian climate changes in central europe on the basis of multi-proxy studies. palaeogeogr. palaeoclimatol. palaeoecol. 461:29-43 krause-dellin d, steinberg c, 1986. cladoceran remains as indicators of lake acidification. hydrobiologia 143:129-134. kupryjanowicz m, 2007. [water level changes in the eemian lakes and peat-bogs in the north podlasie].[article in polish with english abstract]. prz. geol. 55:336-342. kupryjanowicz m, 2008. vegetation and climate of the eemian and early vistulian lakeland in northern podlasie. acta palaeobot. 48:3-130. kupryjanowicz m, ciszek d, mirosław-grabowska j, marciniak b, niska m, 2004. multi-proxy study of eemian deposits from solniki (north-eeastern poland) preliminary results, p. 55. in: proceedings workshop “reconstruction of quaternary palaeoclimate and palaeoenvironments and their abrupt changes”, bialowieza, poland. kupryjanowicz m, ciszek d, mirosław-grabowska j, marciniak b, niska m, 2005. two climatic oscillations during the eemian interglacial preliminary results of multiproxy researches of palaeolake at solniki, ne poland, p. 53-57. polish geological institute special papers n. 16. kurek s, preidl m, 2001. [objaśnienia do arkusza trześcianka szczegółowej mapy geologicznej polski w skali 1: 50,000].[in polish]. central geological archives, pig, warszawa. lindner l, 1992. [czwartorzęd. osady, metody badań, stratygrafia].[in polish]. wyd. pae. mangerud j, 1989. correlation of the eemian and the weichselian with deep sea oxygen isotope stratigraphy. quatern. int. 3-4:1-4. mamakowa k, 1989. late middle polish glaciation, eemian and early vistulian vegetation at imbramowice near wrocław and the pollen stratigraphy of this part of the pleistocene in poland. acta palaeobot. 29:11-176. marks l, 2002. last glacial maximum in poland. quaternary sci. rev. 21:103-110. mirosław-grabowska j, niska m, 2005. isotopic and cladocera records of climate changes of early eemian at besiekierz (central poland). geol. quart. 49:67-74. mirosław-grabowska j, niska m, 2007a. isotope and cladocera data and interpretation from the eemian optimum and postoptimum deposits, kaliska palaeolake (central poland). quatern. int. 175:155-167. mirosław-grabowska j, niska m, 2007b. reconstruction of environmental conditions of eemian palaeolake at studzieniec (central poland) on the basis of stable isotope and cladocera analyses. quatern. int. 162-163:195-204. mirosław-grabowska j, niska m, sienkiewicz e, 2009. evolution of the palaeolake at ruszkówek (central poland) during the eemian interglacial based on isotopic, cladoceran and diatom data. j. paleolimnol. 42:467 481. mirosław-grabowska j, niska m, kupryjanowicz m, 2015. reaction of lake environment on the climatic cooling transition from the eemian interglacial to early vistulian on the basis of solniki palaeolake sediments (ne poland). quatern. int. 386:158-170. mirosław-grabowska j, niska m, roman m, 2016. long (mis 5e 3) environmental history of a paleolake in central poland recorded in the succession from kubłowo. quatern. int. doi: 10.1016/j.quaint.2016.06.027 mirosław-grabowska j, gąsiorowski m, 2010. changes of water level in the eemian palaeolake at imbramowice (sw poland) based on isotopic and cladoceran data. quaternary res. 73:143-150. musiał a, 1992. studium rzeźby glacjalnej północnego podlasia (summary: the study of the glacial sculpture in north podlasie). rozprawy uniwersytetu warszawskiego, 403. nalepka d, walanus a, 2003. data processing in pollen analysis. acta palaeobot. 43:125-134. nauwerck a, 1991. the history of the genus eubosmina in lake mondsee (upper austria). hydrobiologia 225:87-103. nazarova l, 2012. chironomids in quaternary permafrost deposits in the siberian arctic. chironomus newsl. chironomid res. 25:39-42. nevalainen l, luoto tp and sarmaja-korjonen k, 2008. late holocene water-level changes in lake iso lehmälampi, southern finland, reflected in subfossil cladocerans and chironomids. stud. quatern. 25:33-42. nevalainen l, luoto tp, kultti s, sarmaja-korjonen, k 2013. spatio-temporal distribution of sedimentary cladocera (crustacea: branchiopoda) in relation to climate. journal of biogeography 40: 1548-1559. nevalainen l, luoto, tp, 2016. relationship between cladoceran (crustacea) functional diversity and lake trophic gradients. funct. ecol doi:10.1111/1365-2435.12737 niska m, 2008. [interpretacja zmian środowiska jeziornego w interglacjale eemskim na podstawie analizy kopalnych cladocera].[phd thesis in polish]. pomeranian university in słupsk. niska m, 2012. fossil cladocera remains in the eemian sediments preservation, frequency and dominant species. stud. quatern. 29:3143. niska m, 2015. history of the development of eemian interglacial lakes on the basis of cladocera subfossil analysis (central and eastern poland). limnol. rev. 15:85-94. niska m, roman m, 2014. development of the kublowo palaeno n c om me rci al us e o nly eemian/vistulian development of a polish paleolake 219 olake, central poland, during the eemian interglacial as against subfossil cladocera analysis preliminary results. stud. quatern. 31:31-37. niska m, kołodziej a, 2015. history of development and terrestrialization of lake starowlany in the eemian interglacial based on cladocera analysis (sokółka hills, ne poland). stud. quatern. 32:43-52. novakova k, 2005. cladoceran fauna from the archeological site vladar in czech republic. stud. quatern. 22: 51-54. pawłowski d, 2011. evolution of an eemian lake based on cladocera analysis (konin area, central poland). acta geol. polon. 61:441-450. pisias ng, martinson dg, moore tc, shackleton nj, prell w, hays j, boden g, 1984. high resolution stratigraphic correlation of benthic oxygen isotopic records spanning the last 300 000 year. marine geol. 56:119-136. poulsen e, 1944. entomostraca from a late-glacial lacustrine deposit at næstved, denmark. meddelelser fra dansk geologisk forening 10:405-416. robertsson a-m, 2000. the eemian interglacial in sweden, and comparison with finland. geologie en mijnbouw. neth. j. geosci. 79:325-333. rybak ji, błędzki la, 2010. [słodkowodne skorupiaki planktonowe].[book in polish]. university of warsaw: 368 pp. gandouin e, ponel p, andrieu-ponel v, franqu e, de beaulieu jl, reille m, guiter f, brulhet j, lallier-vergès e, keravis d, von grafenstein u, veres d, 2007. past environment and climate changes at the last interglacial/glacial transition (les échets, france) inferred from subfossil chironomids (insecta). geoscience 339:337-346. sarmaja-korjonen k, alhonen p, 1999. cladoceran and diatom evidence of lake-level fluctuations from a finnish lake and the effect of aquatic-moss layers on microfossil assemblages. j. paleolim. 22:277-290. shackleton nj, sánchez-goñi mf, pailler d, lancelot y, 2003. marine isotope substage 5e and the eemian interglacial. global planet. change 3:151-155. shackleton nj, opdyke nd, 1976. oxygen-isotope and paleomagnetic stratigraphy of pacific core v28-239. late pleistocene to latest pleistocene. geol. soc. am. mem. 145. available from: http://memoirs.gsapubs.org/content/145/449. full.pdf shumate bc, schelske cl, crisman tl, kenney wf, 2002. response of the cladoceran community to trophic state change in lake apopka, florida. j. paleolim. 27:71-77. szeroczyńska k, 1985. [cladocera jako wskaźnik ekologiczny w późnoczwartorzędowych osadach jeziornych polski północnej].[article in polish]. acta palaeontol. polon. 30:3-69. szeroczyńska, k. 1998a. [wioślarki (cladocera, crustacea) jako źródło informacji w badaniach osadów jeziornych].[article in polish]. stud. geol. polon. 112:9-28. szeroczyńska k, 1998b. palaeolimnological investigations in poland based on cladocera (crustacea). palaeogeogr. palaeoclimatol. palaeoecol. 140:335-345. szeroczyńska k, 1991. [wioślarki (cladocera) w osadach jeziora skrzetuszewskiego], p. 122-128. in: k. tobolski (ed.), [wstęp do paleoekologii lednickiego parku krajobrazowego]. [book in polish]. biblioteka studiów lednickich. szeroczyńska k, 2002. human impact on lakes recorded in the remains of cladocera (crustacea). quatern. int. 95-96:165-174. szeroczyńska k, 2006. the significance of subfossil cladocera in stratigraphy of late glacial and holocene. stud. quatern. 23:37-45. szeroczyńska k, sarmaja-korjonen k, 2007. atlas of subfossil cladocera from central and northern europe. friends of lower vistula society, świecie: 87 pp. szeroczyńska k, zawisza e, 2007. [paleolimnologia historia rozwoju jezior w polsce w świetle badań fauny wioślarek].[article in polish]. stud. limnol. telmatol. 1: 51-59. tsukada m, 1972. the history of lake nojiri, japan. trans. conn. acad. art. sci. 44:339-365. whiteside mc, 1970. danish chydorid cladocera: modern ecology and core studies. ecol. monogr. 40:79-118. whiteside mc, swindoll mr, 1988. guidelines andlimitations to cladoceran paleoecological interpretations. palaeogeogr. palaeoclimatol. palaeoecol 62:405-412.no n c om me rci al us e o nly layout 1 introduction limnology, as a scientific discipline, was established by françois-alphonse forel in his chief work “le léman: monographie limnologique”, published in three volumes between 1892 and 1904 (warwick and bertola, 2014). this new branch of natural sciences was intended by his founder as “the oceanography of lakes” and represented a holistic approach to the study of biological, chemical, physical and sedimentary aspects of lakes. the discipline very soon attracted wide interest among scientists, and the “societas internationalis limnologiae” (sil, the international society of limnology) was founded one century ago, in 1922, followed in italy by the establishment of the associazione italiana di oceanologia e limnologia (aiol, the italian association of oceanology and limnology) in 1972. studies on the biological aspects of sicilian inland waters started to be published concurrently with the publication of forel’s monograph, and microscopic organisms were the protagonists of some pioneering investigations. at the end of the xix century, it was probably the technical progress of the microscope, and the charming world that this instrument revealed, that attracted several investigators toward the study and description of this microworld. today we know that, among aquatic organisms, microscopic ones (i.e., according to fontaneto and brodie, 2011, in this paper we consider as “microscopic” those organisms belonging to taxa with an average body length of less than 2 mm) play an important role as ecosystem propellers, since they govern primary production, biogeochemical cycles, matter circulation, and are energy conveyors towards larger organisms (e.g., naselli-flores and padisák, 2022). the first investigations on sicilian freshwater microorganisms were those of antonino borzì, who was a pioneer of algological studies, and extensively studied the microalgal flora (including cyanobacteria) of sicily (a list of his contributions is available in borzì, 1996; an anastatic reprint of his work “studi algologici”, originally published in two volumes in 1883 and 1895). a few years later, moniez (1889) reported on the microfauna of some sicilian water bodies and provided the first description of the zooplankton of the “biviere di lentini”, a lake reclaimed in the 1920s to fight malaria (consoli, 1928). in moniez’s paper, the following recommendation appears: “nous ne saurions trop recommander aux lecteurs de la feuille de recueillir avec soin dans les étangs, les lacs et les cours d’eau, les petits animaux et spécialement les crustacés entomostracés, surtout dans les régions montagneuses et méridionales”; the sentence gives us an idea of the urgency and interest, already felt in the second half of the nineteenth century, to investigate the diversity of micro-organisms that populated inland waters in remote and isolated areas. after moniez’s (1889) work, it took more than 35 years for a second short report on the zooplankton of an alkaline lake (the “bagno di venere” or “specchio dell’acqua”) in pantelleria, a volcanic island in the sicilian channel halfway between africa and europe (brehm, 1926). this was followed, a few years later, by an accurate morphological description of a population of arctodiaptomus salinus (daday, 1885) in the endorheic, saline lake pergusa (baldi, 1929). it is worth underlining that the author of this work was edgardo baldi, the first director of the "istituto italiano di idrobiologia dr. marco de marchi" and one of the most renowned limnologists of all time (see https://limnology.org/notable-limnologists/edgardo-baldi/). regardless of this paucity of previous research, limnobiological studies in sicily are in some way related to one of the finest scientists of the last century who tightly linked his name to limnology: george evelyn hutchinson. hutchinson was in palermo at the end of the ‘50s of the last century, and during his stay, he visited the sanctuary of santa rosalia, where he was attracted article a century of research on micro-organisms from the inland waters of the largest mediterranean island federico marrone, luigi naselli-flores department of biological, chemical and pharmaceutical sciences and technologies (stebicef), university of palermo, palermo, italy abstract the first studies on the micro-organisms inhabiting sicilian inland waters date back to the middle of the xix century. however, these were based on single samples and mainly addressed at compiling faunistic and floristic inventories. it was in the first decades of the xx century that the first methodical studies were performed, which focussed on assessing microbial diversity in saline and hypersaline inland waters. studies on plankton dynamics in ponds and reservoirs of the island started at the beginning of the 1980s and, since the end of the 1990s, temporary waters have also been intensively sampled, especially as regards phytoplankton and micro-crustaceans. these intensified sampling efforts contributed to increasing our knowledge of the composition, structure and functioning of the planktic compartment. on the contrary, studies on benthic microflora and fauna are still numerically scarce and mostly based on occasional collections. also, running waters have received little attention and the methodical analysis of their micro-organisms is still in its infancy. no nco mm er cia l u se on ly federico marrone and luigi naselli-flores36 by a nearby small temporary pond. this became the scene of one of his most famous papers (hutchinson, 1959), and probably inspired another seminal paper addressed at solving the so-called “paradox of the plankton” (hutchinson, 1961). in addition, part of the second volume of hutchinson’s monumental work "a treatise on limnology" (hutchinson, 1967), which deals with the analysis of limnoplankton, was written while he was a guest at the institute of zoology of the university of palermo. more than one century has passed since the first studies on micro-organisms inhabiting sicilian inland waters, and our knowledge of their diversity and distribution has greatly increased. updated checklists for some taxa (e.g., rotifers, turbellarians, some crustacean groups) have been recently published in the frame of the new checklist of the italian fauna (bologna et al., 2022), and the occurrence localities of some taxa are reported in the frame of the project ckmap (ruffo and stoch, 2005). however, many gaps still exist, and the complexity related to the analysis of distribution patterns of these organisms is further enhanced by the ongoing climate change; according to woolway and maberly (2020), the velocity of climate change could increase more rapidly than the dispersal velocity of most aquatic species, which are particularly vulnerable to increasing water temperature. this pattern, if confirmed, would reduce the species’ ability to disperse in cooler areas, and thus increase the risk of extinction. as mentioned above, the study of sicilian inland waters is closely related to some of the finest scientists of the last century, who tightly linked their names to limnology and ecology. to pay tribute to these scientists, to celebrate the 50th anniversary of the foundation of the italian association of limnology and oceanography (aiol), and to set a baseline as regards the knowledge available to date, in this paper we try to: i) briefly summarise the work carried out on the micro-organisms of the sicilian inland waters in the last century, and, ii) outline the knowledge gaps that still need to be filled. sicily and its inland waters studying aquatic micro-organisms inhabiting the inland waters of an island located in the middle of the mediterranean sea and characterised by a semi-arid climate could appear, at first sight, somewhat misplaced. however, from a geographical point of view, sicily is part of the southern european border, and an important biogeographical crossroad between subtropical and temperate areas, and the eastern and western part of the mediterranean basin; accordingly, the biological study of sicilian inland waters might allow to better understand the dispersal and distribution patterns of several mediterranean species (e.g., marrone et al., 2009). sicily is the largest island (~25,700 km2) of the mediterranean sea. sicilian physiography is dominated by a hilly landscape (61%) and some mountain ranges (25%, including mount etna, the highest european volcano), with the remaining surface (14%) characterized by lowlands. the climate is typically mediterranean, with an out-of-phase seasonality of rainfall and air temperature that determines the alternance of rainy and cool (winter) and dry and warm (summer) periods. according to the precipitation patterns in the different parts of the island, the climate ranges from semiarid to temperate-dry, with temperate-humid and humid zones limited to the highest mountain ranges (naselli-flores, 2010). due to its insular and climatic conditions, sicily is characterized by a drainage network formed by numerous short, torrent-like, temporary rivers, and a few permanent rivers with a discharge greatly varying between the rainy and the dry periods. a few small, permanent water bodies, and thousands of mediterranean temporary ponds are scattered throughout the island (naselli-flores and marrone, 2019). in addition, several thousands of small agriculture reservoirs have been built in the hilly and lowland parts of the islands in the last 70 years, along with about 30 dam reservoirs with stored volumes ranging between 5 and 150 106 m3 (naselli-flores and lugliè, 2014). the geology of the island strongly conditions the chemical and physical characteristics of these water bodies. an extremely simplified geological division of the island results in two large zones. the northern zone consists of numerous stratigraphic structural units in overthrust, mainly dolomitic and carbonate rocks, and metamorphic rocks. the southern zone is for the most part comprised of carbonate rocks which, in the central part of the island, after the tectonic phase of the middle miocene, were covered by sediments and evaporites, with widespread outcrops of sulphate and gypsum rocks. as a consequence, the ponds and wetlands located in the southern zone often have high conductivity values, generally above 2,000 µs cm-1 (athalassohaline waters). in addition, some saline and brackish lakes and wetlands are located in the plains distributed along the coasts, influenced by the nearby seawater (thalassohaline waters). therefore, the permanent freshwaters are mostly located in the northern mountain chain, at altitudes higher than 900 m a.s.l., and on the hyblean plateau in the south-eastern part of the island. on the whole, a great variety of inland waters are present on the island, from small rockpools to large dam reservoirs, and they include also some relict aquatic environments such as the small, relict peat bogs located in the northern mountain range. the current state of knowledge on micro-organisms of sicilian inland waters in the following paragraphs, we review the state-ofthe-art of some groups of micro-organisms inhabiting a variety of sicilian continental aquatic ecosystems. the review also includes micro-organisms forming large colonies, such as those belonging to the groups of bryno nco mm er cia l u se on ly a century of research on micro-organisms from the inland waters of the largest mediterranean island 37 ozoans and sponges. the papers were collected by searching the internet (google) and using the keywords “sicilian freshwaters” and “sicilian inland waters”, and both the common and taxonomic names of the group to which micro-organisms belong (e.g., “water mites in sicilian inland waters” and “hydrachnidia in sicilian freshwaters”). figure 1 shows the trend in the number of publications on the micro-organisms of sicilian inland waters over the course of more than a century. until the end of the 1970s, most of the contributions were directed to the study of the microbial community of saline and hypersaline lakes. conversely, freshwater micro-organisms have been intensively studied in the last 30 years, especially as regards figure 1. upper panel, the temporal trend of the number of publications on micro-organisms in sicilian inland waters. lower panel, the number of papers containing information on the different groups of micro-organisms. the total number of papers per decade in the upper panel can be higher than the total number of papers appearing in the lower panel since several papers contain information on more than one biological group of micro-organisms. no nco mm er cia l u se on ly federico marrone and luigi naselli-flores38 phytoplankton and microcrustaceans. the publications addressed at this latter group have been continuously increasing in the analyzed period. bacteria and archaea both endorheic and coastal saline lakes have been repeatedly investigated for their bacterial communities since the beginning of the 1930s. lake pergusa is the largest sicilian natural lake, and is located on the evaporite outcrops covering the center of the island. it was an early subject of research due to its cyclic reddish water blooms, due to purple sulphur bacteria that periodically bloomed in its hypersaline waters (brunelli and maldura, 1929; forti, 1932; 1933a; 1933b). further investigations on this phenomenon were carried out in the late 1970s (faranda et al., 1977; genovese et al., 1977; bruni and pulicanò, 1978) and at the beginning of this century (de francesco et al., 2002). urbanization and increased water demand for agriculture in the surroundings of the lake promoted intensive groundwater extraction (canova et al., 2011), which caused a strong reduction in the water level at the beginning of this century, threatening its existence. to counteract this tendency, since 2005, the lake has been refilled using fresh waters from dam-reservoirs with much lower conductivity values (lake ancipa, ~200 µs cm-1). this induced a quite drastic rearrangement of the biota in lake pergusa as discussed in barone et al. (2008) and naselli-flores et al. (2016), and caused the disappearance of the purple sulphur bacteria cyclic blooms. lake faro, a meromictic and quite deep coastal lake was the theatre of several studies on its peculiar bacterial community. since the second half of the 1950s, and for more than half a century, the lake has attracted the attention of scientists due to the stable presence of hydrogen sulphide in the monimolimnion and of a layer of red water at depths between 9 and 12 meters (genovese, 1957; 1961; 1963), caused by phototrophic sulphur bacteria (genovese et al., 1958; 1962; genovese and machi, 1963; trüper and genovese 1968; sorokin and donato 1975; bruni et al., 1978; acosta-pomar et al., 1988; maugeri et al., 2001; saccà et al., 2008; gugliandolo et al., 2011; lentini et al., 2012). isolated investigations on the bacterial community of tindari ponds (zaccone et al., 2000; caruso et al., 2005) and river alcantara (guglielmo et al., 2013) were carried out in the first years of this century. besides the pioneering work carried out by antonino borzì on benthic cyanobacteria (see borzì, 1996, for a detailed list of papers), no additional information is available in the literature on these micro-organisms other than the recent paper by russo et al. (2021), who studied the diversity of cyanobacteria colonizing the surface of an urban fountain in palermo by combining microscopic analysis and molecular tools. conversely, the distribution and abundance of planktic cyanobacteria in sicilian water bodies (barone and naselli-flores, 1989) was first systematically assessed in a limnological investigation aimed at evaluating the trophic state of all the sicilian lakes and reservoirs with a surface area >0.2 km2 (calvo et al., 1993). since cyanobacteria are commonly found in the plankton of lakes and ponds, information on this group of micro-organisms is available in the papers on phytoplankton of sicilian inland waters (see next paragraph). phytoplankton and microbenthic algae faranda (1977) compiled the first inventory of the sicilian water bodies suitable for aquaculture and investigated their main chemical, physical and biological (fish) characteristics. the paper also contains some information on phytoand zooplankton, but it is largely incomplete, since this information is only available for 7 out of the 22 studied water bodies, and just a few species are listed. phytoplankton, including the “pico” fraction, in a complex system of brackish and saline inland waters (capo peloro lagoon system, which also includes lake faro) was investigated in detail, over a multi-year period, by platt et al. (1985); bruni et al. (1990); giuffrè (1991; 2011); giacobbe and maimone (1994); gangemi (2000); gangemi et al. (2006); giuffrè and pezzani (2005). the first studies on the diversity, abundance, and annual dynamics of freshwater phytoplankton in different sicilian man-made lakes are those by barone (1983; 1985) and barone et al. (1989). these works were prodromal for a more comprehensive investigation of plankton from sicilian reservoirs (calvo et al., 1993; barone and naselli-flores, 1990; barone et al., 1991), and entailed a greater effort in the analysis of phytoplankton assemblages in sicilian permanent waters, based on weekly samplings, which followed the phytoplankton generation times (barone and naselli-flores, 1994; 1995; naselliflores and barone, 1994; 1995; 1998; 2003). this sampling frequency, carried out for almost twenty years in the dam-reservoir “lake arancio” (naselli-flores, 2014), allowed to clarify several ecological aspects of the phytoplankton dynamics in the mediterranean dam-reservoirs, and to establish a link between phytoplankton dynamics and the peculiar morphological and hydraulic characteristics of these water bodies (naselli-flores, 1996; 1998; 1999; 2000; 2003; 2010; naselli-flores and barone, 2007; 2011). furthermore, the importance of the major role exerted on plankton dynamics by the “filling-emptying” dynamics in mediterranean reservoir, compared to the “stratification-circulation” paradigm in temperate lakes (e.g., sommer, 1987) was highlighted (naselli-flores, 2003), as well as the usefulness of a functional classification of planktic organisms based on traits (barone and naselli-flores, 1990; naselli-flores and barone, 2000). the “filling-emptying” dynamics in sicilian reservoirs no nco mm er cia l u se on ly a century of research on micro-organisms from the inland waters of the largest mediterranean island 39 also resulted in favouring the growth of toxic cyanobacteria like microcystis spp., with the formation of long-lasting blooms in those years characterized by wider fluctuations in the water level (naselli-flores et al., 2007). a partial inventory of freshwater phytoplankton (including cyanobacteria) in sicilian inland water >0.2 km2 can be found in barone (2003). phytoplankton studies in permanent and temporary natural ponds in sicily were mainly addressed at understanding the effects of climate change on phytoplankton composition and dynamics (barone et al., 2010), how regional and local factors govern the phytoplankton assembly (naselli-flores and barone, 2005; 2012; jeppesen et al., 2015) and how passive dispersal can influence phytoplankton distribution in natural and temporary waters (incagnone et al., 2015; naselli-flores et al., 2016). factors influencing the seasonal distribution of mixotrophic phytoplankton were also studied over a multi-year (6.5) period (naselli-flores and barone, 2019). moreover, sicilian permanent and temporary ponds host a peculiar microflora and several rare species (naselliflores and barone, 2000; 2012; barone et al., 2006; madonia et al., 2006). in particular, the ponds located on the northern mountain range host a relict microflora typical of more temperate aquatic ecosystems (barone et al., 1989; barone, 2003). a complete account of the biodiversity of these ponds deserves further investigations, as shown in a recent paper by pang and van de vijver (2021), who investigated the chrysophycean stomatocyst flora in some small pools on the hyblean mountains in south-eastern sicily. microbenthic eukaryotic algae have seldom been investigated in sicilian inland waters and, apart from an in-depth study on the benthic diatoms of the simeto hydrographic basin (finocchiaro et al., 2009; 2011), the most extensive river network of the island, only a few further contributions are available on benthic diatoms in a coastal meromictic lake, and in freshwater and saline rivers (giuffrè, 2000; mannino, 2001; 2007; campisi et al., 2020). rotifers free swimming rotifers of the class monogononta commonly inhabit freshwaters and represent an important component of the zooplankton. moniez (1889) was the first who studied rotifers in sicilian inland waters: in his paper this author writes “on n'a encore rien publié, que je sache, sur la faune des eaux douces de la sicile”. this first contribution provided a list of eight species from a shallow lake in the eastern part of the island, the biviere di lentini. the original water body was drained at the end of the 1920s to fight malaria (consoli, 1928) but at the end of the last century, a new dam-reservoir with the same name was built in a nearby location. among the eight species found by moniez in the “old” biviere di lentini, anuraea longispina kellicot, 1879, is listed. this species, whose accepted name is kellicottia longispina (kellicot, 1879), is considered a cold stenothermic taxon (jersabek and bolortsetseg, 2010) and thus represent an unusual record for the island. however, naselli-flores (unpublished) in 1987 found a single specimen of this species in a single zooplankton winter sample collected in the biviere di cesarò, a sicilian mountain pond located at 1274 m a.s.l. (naselli-flores, unpublished). due to the occasional nature of the finding, it was decided not to include the record in a paper on the zooplankton of the biviere di cesarò (cf. naselli-flores and barone, 1991) and no further records for this species are to date known for the island. the “unusual” report of kellicottia longispina was also noted by berzins (1954), who, in a fruitless attempt to find it, collected and examined some samples in two water bodies which were located close to the, at that time disappeared, biviere di lentini. in his paper, berzins accurately listed and described 37 species of rotifers (mainly belonging to the class monogononta) along with six species of cyclopoid copepods and eight cladocerans. after the publication of the paper by berzins, the monogonont fauna in sicilian inland waters was seldom investigated, and no information is available on bdelloid rotifers. a further list of rotifers, collected in a series of seasonal surveys on the plankton of sicilian lakes and reservoirs, was published almost 40 years after the paper by berzins (calvo et al., 1993), whereas the first data on rotifers in lotic ecosystems were recently provided by guglielmo et al. (2013) and rodriguez et al. (2013) who studied the rotifer assemblage of river alcantara. an extensive review of rotifers occurring in sicilian inland and marine waters, including geo-referenced occurrence localities, is reported in ferrari et al. (2023). microcrustaceans information on microcrustaceans in sicilian inland waters is scattered in several papers. these include the first observations by moniez (1889) on the zooplankton of the biviere di lentini, and by brehm (1926) and baldi (1929) on the calanoid copepod arctodiaptomus salinus (daday, 1885). cannicci (1939) reports on the plankton collected in a small, hypersaline pond on the island of salina (aeolian archipelago) and lists, among others, the occurrence of a harpacticoid copepod, mesochra aestuarii gurney, 1921. the first inventory of microcrustaceans from sicilian freshwaters was provided by margaritora et al. (1982), who studied in detail the entomostracan fauna collected in some dam-reservoirs, and in several permanent and temporary ponds. in the same period, the copepods of sicilian groundwaters were investigated in detail (e.g., pesce, 1988, 1994; pesce and galassi, 1987; pesce et al., 1987, 1988). these sampling efforts were further integrated by a study, carried out with seasonal frequency, on the trophic state of all sicilian lakes and reservoirs with a surface area larger than 0.2 km2 (calvo et al., 1993). in no nco mm er cia l u se on ly federico marrone and luigi naselli-flores40 that occasion, crustacean zooplankton was also collected and identified and further records were added to the knowledge on the distribution and diversity of microcrustacean fauna in sicilian inland waters (naselli-flores et al., 1998; alfonso et al., 2010). moreover, long-term weekly samplings allowed to clarify the ecological relationships between seasonal water-level fluctuations and cladoceran dynamics in a mediterranean dam-reservoir (naselli-flores and barone, 1997). however, the picture was far from exhaustive as demonstrated by the numerous new records of microcrustaceans (calanoid, cyclopoid, and harpacticoid copepods, cladocerans and ostracods) obtained in the new century through an intensified sampling activity in sicilian groundwaters (e.g., cottarelli et al., 2012; bruno et al., 2018), and in some hundreds of temporary ponds (e.g., marrone, 2003, 2006; marrone and naselli-flores, 2004, 2005; marrone and vecchioni, 2020; marrone et al., 2005; 2006a, b, 2020; castelli et al., 2006; pieri et al., 2006, 2020; vecchioni et al., 2019a, 2021a; marchese et al. 2022). these studies have contributed to increasing knowledge of the biodiversity of sicilian inland waters, making the island one of the best-studied territories in the mediterranean basin. furthermore, they allowed: to clarify taxonomic problems and biogeographical issues related to micro-crustaceans distribution in the mediterranean inland waters (marrone et al., 2009; 2017; mazzini et al., 2017; alfonso et al., 2021; vecchioni et al., 2021b; kotov et al., 2022); to better understand speciation in copepods (marrone et al., 2010; 2013; 2016); to analyze species extinction, coexistence and replacement (vecchioni et al., 2019b); and to account for the establishment of non-indigenous species (marrone and naselli-flores, 2015; naselli-flores and marrone, 2019). other micro-organisms moniez's pioneering work (moniez, 1889) reports a number of aquatic micro-organisms (e.g., protozoa, hydrozoa, water mites) that were not further investigated or received little attention in the following studies on the biota of sicilian inland water. apart from some faunal and systematics works on water mites (gerecke et al., 2014 and literature therein) and tardigrades (lisi, 2015 and literature therein), knowledge of other groups of aquatic or semiaquatic micro-organisms is incomplete and fragmentary. heterotrophic protozoa (e.g., heterotrophic microand nanoflagellates as well as ciliophora, both peritrichida and tintinnida, and amoeba-like protozoans and testate amoebae belonging to arcellinida) are often observed in live samples of plankton but taxonomic and ecological studies on this group of organisms for sicilian inland waters are lacking. in the same way, freshwater turbellarian flatworms occurring in the island are poorly known (but see de vries, 1985 and literature therein, and lázaro and riutort, 2013 and literature therein); colomba and vinciguerra (1980) provided an account of free-living freshwater nematodes in the river anapo, whereas vecchioni et al. (2021c) reported of an ectosymbiotic platyhelminth on freshwater crustaceans from an aquaculture farm. as regard freshwater hydrozoans, after the finding by moniez (1889) of hydra fusca linnaeus, 1767 (synonym of h. oligactis pallas, 1766), the only published report in the last 130 years for sicilian inland water is related to the finding of hydra viridissima pallas, 1766 and h. oligactis in the river oreto near palermo (riggio, 1978), and of a population of craspedacusta sowerbii lankester, 1880 in a small concrete-reservoir in the city of palermo (schifani et al., 2019). bryozoans were reported by riggio (1978) and studied by taticchi et al. (2011) who found 7 species by sampling 4 dam-reservoirs and the river anapo. the first record of a freshwater sponge in sicily date back to 1961, when a population of ephydatia fluviatilis (linnaeus, 1759) was sampled in a fountain close to the greek theatre of syracuse (sarà, 1961). riggio (1978) found this species in the river oreto, and manconi and pronzato (2016) reported of a population in a temporary canal in north-western sicily. these further findings suggest that the sponge might be quite widespread on the island. conclusions although sicily (and its surrounding archipelagos) is probably among the best studied italian regions regarding the biodiversity of its inland waters, new and interesting information emerge whenever new environments are studied or when understudied taxonomic groups are considered. as an example, marrone et al. (2022) highlighted some cases of apparent counter-intuitive distribution patterns and ecological preferences for sicilian calanoid copepods, possibly to be ascribed to niche displacement or inaccurate taxonomy. dam-reservoirs, and permanent and temporary ponds have been studied over multiannual periods as regards their plankton, but the information on several groups of micro-organisms is still far from being complete. moreover, systematic surveys aimed at assessing the biodiversity of micro-organisms in permanent and temporary rivers are highly desirable. new methods such as the implementation of edna surveys can provide valuable information and represent an interesting opportunity to assess biodiversity in sicily as revealed by a study carried out by hupało et al. (2022) in a sicilian stream. however, it should also be stressed that the method has shown some clear downsides (e.g., roussel et al., 2015). among these, the paucity and poorly representativeness of the available reference libraries for dna sequences of the micro-organisms of the mediterranean area might constitute a potential pitfall for a sound species no nco mm er cia l u se on ly a century of research on micro-organisms from the inland waters of the largest mediterranean island 41 identification and assessment, so that before successfully becoming an efficient tool to investigate and monitor aquatic species diversity, edna surveys must be preceded by accurate studies and characterization of the local biological diversity based on an integrative approach. in order to get a representative and sound assessment of the biodiversity of micro-organisms inhabiting sicilian inland water bodies, the formation and financial support of a new generation of specialists is really mandatory. as often stressed (e.g., engel et al., 2021), the taxonomic impediment, and the consequent “linnean shortfall” (brown and lomolino, 1998), is due to a lack of taxonomists, not to a lack of available techniques! corresponding author: luigi naselli-flores, department of biological, chemical and pharmaceutical sciences and technologies (stebicef), university of palermo, palermo, italy. e-mail: luigi.naselli@unipa.it. authors’ contributions: the authors contributed equally to the manuscript. conflict of interest: the authors declare no potential conflict of interest. funding: none. availability of data and materials: all data generated or analyzed during this study are included in this published article. key words: freshwater biota, sicily, microalgae, microcrustaceans, rotifers, microinvertebrates. received: 24 novemberr 2022. accepted: 5 december 2022. publisher’s note: all claims expressed in this article are solely those of the authors and do not necessarily represent those of their affiliated organizations, or those of the publisher, the editors and the reviewers. any product that may be evaluated in this article or claim that may be made by its manufacturer is not guaranteed or endorsed by the publisher. ©copyright: the author(s), 2022 licensee pagepress, italy advances in oceanography and limnology, 2022; 13:11033 doi: 10.4081/aiol.2022.11033 this work is licensed under a creative commons attribution noncommercial 4.0 international license (cc by-nc 4.0). references acosta pomar l, bruni v, decembrini f, giuffrè g, maugeri tl, 1988. distribution and activity of picophytoplankton in a brackish environment. progress in oceanography 21:129-138. alfonso g, belmonte g, marrone f, naselli-flores l, 2010. does lake age affect zooplankton diversity in mediterranean lakes and reservoirs? a case study from southern italy. hydrobiologia 653:149-164. alfonso g, stoch f, marrone f, 2021. an annotated checklist and bibliography of the diaptomidae (copepoda calanoida) of italy, corsica and the maltese islands. journal of limnology 80:2019. baldi e, 1929. diaptomus salinus daday nel lago di pergusa (sicilia). bollettino di pesca, piscicultura e idrobiologia 5:827846. barone r, 1983. fitoplancton e stato trofico del lago poma. atti 5° congresso alol:177-187. barone r, 1985. fitoplancton e stato trofico del lago di piana degli albanesi. bollettino dell'accademia gioenia di scienze naturali, catania 18:485-498. barone r, 2003. a critical inventory of freshwater phytoplankton in sicilian lakes. bocconea 16:355-365. barone r, castelli g, naselli-flores l, 2010. red sky at night cyanobacteria delight: the role of climate in structuring phytoplankton assemblage in a shallow, mediterranean lake (biviere di gela, southeastern sicily). hydrobiologia 639:43-53. barone r, marrone f, naselli-flores l, 2006. first record of cyanophora paradoxa koršikov (glaucocystophyta) in italy. naturalista siciliano, series iv, 30:97-106. barone r, naselli-flores l, 1989. cianoficee planctoniche dei laghi siciliani: osservazioni preliminari. giornale botanico italiano 123:74. barone r, naselli-flores l, 1990. osservazioni sulle comunità planctoniche dei laghi artificiali siciliani: relazioni diversità biomassa. naturalista siciliano, series iv, 14:35-48. barone r, naselli-flores l, 1994. phytoplankton dynamics in a shallow, hypertrophic reservoir (lake arancio, sicily). hydrobiologia 289:199-214. barone r, naselli-flores l, 1995. strategie del fitoplancton e modificazioni nella struttura delle comunità: un caso di studio. giornale botanico italiano 129:189-197. barone r, naselli-flores l, calvo s, 1989. fitoplancton e livello trofico del lago biviere di cesarò. naturalista siciliano, series iv, 13:1-22. barone r, naselli-flores l, calvo s, 1991. plankton communities in the artificial lakes of sicily (italy). verhandlungen der internationalen vereinigung für theoretische und angewandte limnologie 24:1409-1414. barone r, termine r, massa b, 2008. note su un bloom di prymnesium parvum carter (haptophyta) nel lago di pergusa e sue conseguenze sulla fauna. naturalista siciliano, series iv, 32:187-200. berzins b, 1954. zur rotatorienfauna siziliens. hydrobiologia 6:309-320. bologna ma, bonato l, cianferoni f, minelli a, oliverio m, stoch f, zapparoli m, 2022. towards the new checklist of the italian fauna. biogeographia the journal of integrative biogeography 37:ucl001. borzì a, 1996. studi algologici: saggio di ricerche sulla biologia delle alghe (fascicoli i e ii – tavole i-xxxi). ristampa anastatica a cura di f.m. raimondo and l. naselli-flores. naturama, palermo, italy: 480 pp. brehm v, 1926. die mikrofauna des alkalinischen seen auf pantelleria. in: mertens r. (ed.), zoologische ergebnisse einer reise nach den pelagischen inseln und sizilien. senckenbergiana, band 8, heft 5/6:262-265 brown jh, lomolino mv, 1998. biogeography. 2nd ed. sinauer no nco mm er cia l u se on ly federico marrone and luigi naselli-flores42 associates inc., sunderland, massachusetts, usa: 691 pp. brunelli g, maldura c, 1929. osservazioni sul lago di pergusa. bollettino di pesca, piscicoltura e idrobiologia 4:594-604. bruni v, acosta pomar mlc, giuffrè g,1990. picoplankton and picophytoplankton in a brackish environment (lake of ganzirri messina). — rapp. comm. int. mer médit. 32:137. bruni v, pulicanò g, 1978. nuovi dati sulle condizioni fisicochimiche e microbiologiche del lago di pergusa. atti della società peloritana di scienze fisiche, matematiche e naturali 24:265-280. bruni v, pulicanò g, arco a, 1978. ulteriori ricerche sui batteri del ciclo dello zolfo nel lago di faro. atti della società peloritana di scienze fisiche, matematiche e naturali 24:291-299. bruno mc, cottarelli v, grasso r, latella l, zaupa s, spena mt, 2018. epikarst crustaceans from some italian caves: endemisms and spatial scales. biogeographia – the journal of integrative biogeography 33:1-18. calvo s, barone r, naselli-flores l, fradà orestano c, dongarrà g, lugaro a, genchi g, 1993. limnological studies on lakes and reservoirs of sicily. naturalista siciliano, s. iv, 17:1-292. campisi p, mannino am, venturella g, ravera s, 2020. a first contribution to the cryptogamic flora of “bosco pomieri” (northern sicily). borziana 1:35-51. cannicci g, 1939. prime osservazioni sul plancton di alcuni stagni salmastri mediterranei in rapporto alla salinità. bollettino di pesca, piscicoltura e idrobiologia 15:108-145. caruso g, monticelli l, azzaro m, decembrini f, la ferla r, leonardi m, zaccone r, 2005. dynamics of extracellular enzymatic activities in a shallow mediterranean ecosystem (tindari ponds, sicily). marine and freshwater research 56:173-188. castelli g, marrone f, barone r, naselli-flores l, 2006. crustacean dynamics in two argillotrophic temporary ponds (northwestern sicily, italy). vereinigung für theoretische und angewandte limnologie 29:1265-1270. colomba g, vinciguerra mt, 1980. freshwater nematodes of sicily. i. nematodes of the anapo. animalia 6:89-120. consoli, n, 1928. gl’interventi di piccola bonifica nella lotta contro la malaria in sicilia. rivista sanitaria siciliana 18, francesco sanzo e c., industria tipografica editrice, palermo, italy: 108 pp. cottarelli v, bruno mc, spena mt, grasso r, 2012. studies on subterranean copepods from italy, with descriptions of two new epikarstic species from a cave in sicily. zoological studies 51:556-582. de francesco m, pulicanò g, termine r, bruni v, 2002. red water in the lake of pergusa (sicily, enna). 8th symposium on aquatic microbial ecology (same 8), 25-30 october 2002, taormina, italy. available at: http://www.riserveenna.it/portals/riserveenna/documenti/p ergusa/same8_brunitermine.pdf de vries ej, 1985. the biogeography of the genus dugesia (turbellaria, tricladida, paludicola) in the mediterranean region. journal of biogeography 12:509-518. engel ms, ceríaco lmp, daniel gm, dellapé pm, löbl i, marinov m, et al., 2021. the taxonomic impediment: a shortage of taxonomists, not the lack of technical approaches. zoological journal of the linnean society 193:381–387. faranda f, 1977. primo censimento delle aree destinabili ad acquacoltura in sicilia. atti della società peloritana di scienze fisiche, matematiche e naturali 23:3-114. faranda f, bruni v, de domenico e, 1977. nouvelles observations sur le lac de perguse (sicile). rapport du comité international pour la mer méditerranéenne 24:67-70. ferrari v, gualdi a, bertani i, fontaneto d, kamburska l, karimullah k, marrone f, obertegger u, rossetti g, tiberti r, cancellario t, 2023. a georeferenced dataset of italian occurrence records of the phylum rotifera. journal of limnology (in press). finocchiaro m, torrisi m, ferlito a, 2009: caratterizzazione delle comunità di diatomee bentoniche del bacino idrografico del fiume simeto (sicilia orientale) mediante applicazione dell’indice diatomico di eutrofizzazione/polluzione (epi-d). biologia ambientale 23:53-66. finocchiaro m, torrisi m, ferlito a, dell’uomo a, ector l, 2011. biodiversità algale in sicilia: diatomee rare o notevoli del bacino idrografico del fiume simeto. biogeographia 30:13–29. fontaneto d, brodie j, 2011. why biogeography of micro-organisms? in fontaneto, d. (ed.), biogeography of microscopic organisms. is everything small everywhere? cambridge university press, cambridge, uk. available at: https://www.cambridge.org/core/books/abs/biogeographyof-microscopic-organisms/why-biogeography-of-microorganisms/fc732fcccbb2d4fdebca3619f32314da forti a, 1932. il lago di sangue a pergusa in sicilia e la prima piaga d’egitto. naturalista siciliano 8:63-86 forti a, 1933a. nuove notizie sull'arrossamento delle acque avvenuto nel lago di pergusa in settembre 1932 ed ulteriori considerazioni sui fenomeni di arrossamento in generale. bollettino di pesca, piscicotura e idrobiologia 9:998-1019. forti a, 1933b. il fenomeno del lago di sangue nello stagno di pergusa in sicilia alla metà di settembre 1932. nuovo giornale botanico italiano 40:76-78 gangemi e, 2000. dinoflagellati nocivi nelle fitocenosi di un ecosistema litorale della sicilia orientale (pantano grande lagune di capo peloro me). tesi di dottorato, università di messina. available at: https://opac.bncf.firenze.sbn.it/bncfprod/resource?uri=tsi0101477 gangemi e, giacobbe mg, fraga s, 2006. il complesso delle specie ciguateriche nelle lagune di capo peloro (messina, sicilia). biologia marina mediterranea 13:979-983. genovese s, 1957. ciclo batterico dello zolfo con particolare riferimento agli stagni salmastri. atti della società peloritana di scienze fisiche, matematiche e naturali 3:361-388. genovese s, 1961. sur la présence d’“eau rouge” dans le lac de faro (messine). rapports et procès-verbaux des réunions commission internationale pour l'exploration scientifique de la mer méditerranée 16:255-256. genovese s, 1963. the distribution of h2s in the lake of faro (messina) with particular regard to the presence of "red water". in: oppenheimer e.h. (ed.), symposium on marine microbiology. thomas, springfield, usa:194-204. genovese s, bruni v, faranda f, 1977. sulle attuali condizioni del lago di pergusa. rivista di idrobiologia 16:245-268. genovese s, machi g, 1963. ciclo annuale di osservazioni microbiologiche nel lago di faro. atti della società peloritana di scienze fisiche, matematiche e naturali 9:293-329. no nco mm er cia l u se on ly a century of research on micro-organisms from the inland waters of the largest mediterranean island 43 genovese s, pichinoity f, senez sc, 1958. sui batteri solfatoriduttori del lago di faro (messina). la ricerca scientifica, rivista del consiglio nazionale delle ricerche 28:131-140. genovese s, rigano c, la cava m, 1962. ulteriori osservazioni sulla presenza dell’ “acqua rossa” nel lago di faro. atti della società peloritana di scienze fisiche, matematiche e naturali 8:503-510. gerecke r, marrone f, sorgi g, dossena m, stoch f, 2014. the water mites (acari: hydrachnidia) of the standing waters of corsica, sardinia and sicily: review and new data. italian journal of zoology 81:389-408. giacobbe mg, maimone g, 1994. first report of alexandrinum minutum halim in a mediterranean lagoon. cryptogamie algologie 15:47-52. giuffrè g, 1991. le diatomee planctoniche del lago salmastro meromittico di faro, messina. un anno di osservazioni. bollettino della società adriatica di scienze 72:51-79. giuffrè g, 2000. the diatoms of the faro lagoon (sicily). 16th international diatom symposium, athens & aegean islands, 25 august 1 september 2000: proceedings. giuffrè g, 2011. le diatomee planctoniche del sistema lagunare di capo peloro (messina, sicilia) e la successione annuale del fitoplancton in rapporto con i parametri fisico-chimici nel pantano di ganzirri. naturalista siciliano, series iv, 35:187-214. giuffrè g, pezzani r, 2005. annual nano-microphytoplanton succession in the meromictic brackish lake faro (messinasicily) in relation to physical-chemical conditions. naturalista siciliano, series iv, 29:153-168. gugliandolo c, lentini v, maugeri tl, 2011. distribution and diversity of bacteria in a saline meromictic lake as determined by pcr-dgge of 16s rrna gene fragments. current microbiology 62:159-66. guglielmo l, azzaro f, baviera c, bergamasco a, bissett sn, brugnano c, et al., 2013. multidisciplinary ecological assessment of the alcantara river (sicily, italy) using bioindicators. marine and freshwater research 65:283-305. hupało k, schmidt s, macher th, weiss m, leese f, 2022. fresh insights into mediterranean biodiversity: environmental dna reveals spatio-temporal patterns of stream invertebrate communities on sicily. hydrobiologia 849:155-173. hutchinson ge, 1959. homage to santa rosalia or why are there so many kinds of animals? the american naturalist 93:145-159. hutchinson ge, 1961. the paradox of the plankton. the american naturalist 95:137-145 hutchinson ge, 1967. a treatise on limnology, volume 2. introduction to lake biology and the limnoplankton. john wiley & sons, new york, usa: 1115 pp. incagnone g, marrone f, barone r, robba l, naselli-flores l, 2015. how do freshwater organisms cross the “dry ocean”? a review on passive dispersal and colonization precesses with a special focus on temporary ponds. hydrobiologia 750:103-123. jeppesen e, brucet s, naselli-flores l, papastergiadou e, stefanidis k, nõges t, et al., 2015. ecological impacts of global warming and water abstraction on lakes and reservoirs due to changes in water level and related changes in salinity. hydrobiologia 750:201-227. jersabek cd, bolortsetseg e, 2010. mongolian rotifers (rotifera, monogononta) – a checklist with annotations on global distribution and autoecology. proceedings of the academy of natural sciences of philadelphia 159:119-168. kotov aa, garibian pg, neretina an, marrone f, 2022. a redescription of the mediterranean endemic cladoceran daphnia chevreuxi richard, 1896 (cladocera: daphnidae). zootaxa 5125:205-228. lázaro em, riutort m, 2013. dugesia sicula (platyhelminthes, tricladida): the colonizing success of an asexual planarian. bmc evolutionary biology 13:268. lisi o, 2015. current knowledge on the sicilian tardigrade fauna. biodiversity journal 6:297-304. madonia p, naselli-flores l, parello f, parlato b, viola a, 2006. geological development of a gypsum lake formed at the beginning of the 20th century in central sicily, italy: integration of historical data with modern survey techniques. chemistry and ecology 22:333-347. manconi r, pronzato s, 2016. how to survive and persist in temporary freshwater? adaptive traits of sponges (porifera: spongillida): a review. hydrobiologia 782:11-22. mannino am, 2001. new records from the oreto river (n.w. sicily). phycologia 40:71-72 mannino am, 2007. diatoms for thermal-sulphur waters of “fiume caldo” (north-western sicily). cryptogamie, algologie 28:385-396. marchese m, vecchioni l, bazan g, arculeo m, marrone f, 2022. the inland water copepod fauna of a traditional rural landscape in a mediterranean island (crustacea, copepoda). water 14:2168. margaritora fg, mastrantuono l, crosetti d, lombardi f, 1982. contributo allo studio della fauna ad entomostraci delle acque interne della sicilia. animalia 9:87-102. marrone, f. 2003. on some cladocerans new to sicily (southern italy) (crustacea branchiopoda). naturalista siciliano, series iv, 27:263-270. marrone, f. 2006. the microcrustacean fauna of sicily and the central mediterranean sea area current knowledge and gaps to be filled. polish journal of ecology 54:681-685. marrone f, alfonso g, naselli-flores l, stoch f, 2017. diversity patterns and biogeography of diaptomidae (copepoda, calanoida) in the western palearctic. hydrobiologia 800: 45-60. marrone f, barone r, naselli-flores l, 2005. cladocera (branchiopoda: anomopoda, ctenopoda, and onychopoda) from sicilian inland waters: an updated inventory. crustaceana 78:1025-1039. marrone f, barone r, naselli-flores l, 2006. ecological characterization and cladocerans, calanoid copepods and large branchiopods of temporary ponds in a mediterranean island (sicily, southern italy). chemistry and ecology 22:s181-s190. marrone f, castelli g, barone r, naselli-flores l, 2006. ecology and distribution of calanoid copepods in sicilian inland waters. vereinigung für theoretische und angewandte limnologie 29:2150-2156. marrone f, castelli g, naselli-flores l, 2009. sicilian temporary ponds: an overview of the composition and affinities of their crustacean biota, p. 189-202. in: i. fraga and p. arguimbau (eds.), international conference on mediterranean temporary ponds. proceedings & abstracts. consell insular de menorca. recerca, 14. maó, menorca. no nco mm er cia l u se on ly federico marrone and luigi naselli-flores44 marrone f, fontaneto d, naselli-flores l, 2022. cryptic diversity, niche displacement and our poor understanding of taxonomy and ecology of aquatic micro-organisms. hydrobiologia. available at: https://link.springer.com/article/10.1007/s10750-022-04904-x marrone f, havenstein k, tiedemann r, ketmaier v, 2016. identification and characterization of five polymorphic microsatellite loci in the freshwater copepod hemidiaptomus gurneyi (copepoda: calanoida: diaptomidae). italian journal of zoology 83:146-150. marrone f, lo brutto s, arculeo m, 2010. molecular evidence for the presence of cryptic evolutionary lineages in the freshwater copepod genus hemidiaptomus g.o. sars, 1903 (calanoida, diaptomidae). hydrobiologia 644:115-125. marrone f, lo brutto s, hundsdoerfer ak, arculeo m, 2013. overlooked cryptic endemism in copepods: systematics and natural history of the calanoid subgenus occidodiaptomus borutzky 1991 (copepoda, calanoida, diaptomidae). molecular phylogenetics and evolution 66:190-202. marrone f, naselli-flores l, 2004. first record and morphological features of hemidiaptomus (occidodiaptomus) ingens (gurney, 1909) (copepoda calanoida) in italy. journal of limnology 63:250-255. marrone f, naselli-flores l, 2005. first record of a representative of the subfamily paradiaptominae (copepoda calanoida diaptomidae) in italy: metadiaptomus chevreuxi (guerne and richards, 1894). journal of limnology 64:89-92. marrone f, naselli-flores l. 2015. a review on the animal xenodiversity in sicilian inland waters (italy). advances in oceanography and limnology 6:2–12. marrone f, ortega f, mesquita-joanes f, guerrero f, 2020. on the occurrence of metadiaptomus chevreuxi (calanoida, diaptomidae, paradiaptominae) in the iberian peninsula, with notes on the ecology and distribution of its european populations. water 12:1989. marrone f, vecchioni l, 2020. the genus daphnia in sicily and malta (crustacea, branchiopoda, anomopoda), p. 105-123. in: t. la mantia, e. badalamenti, a. carapezza, p. lo cascio and a. troia (eds.), life on islands. 1. biodiversity in sicily and surrounding islands. studies dedicated to bruno massa. edizioni danaus, palermo. maugeri tl, gugliandolo c, caccamo d, la rosa t, 2001. biomassa e produttività delle comunità microbiche negli ambienti salmastri di ganzirri e faro (messina). biologia marina mediterranea 8:1–10. mazzini i, marrone f, arculeo m, rossetti g, 2017. revision of recent and fossil mixtacandona klie 1938 (ostracoda, candonidae) from italy, with description of a new species. zootaxa 4221: 323-340. moniez e, 1889. note sur la faune des eaux douces de la sicile. feuille des jeunes naturalistes 230: 17-19. naselli-flores l, 1996. the structure of phytoplankton assemblages in sicilian reservoirs: some hypotheses on the driving factors. giornale botanico italiano 130:1004-1007. naselli-flores l, 1998. phytoplankton assemblages in reservoirs: is it chemical or physical constraints which regulate their structure? international review of hydrobiology 83:351–360. naselli-flores l, 1999. limnological aspects of sicilian reservoirs: a comparative, ecosystemic approach, p. 283-311. in: j.g. tundisi and m. straškraba (eds.), theoretical reservoir ecology and its applications. backhuys publishers, leiden, the netherlands. naselli-flores l, 2000. phytoplankton assemblages in twentyone sicilian reservoirs: relationship between species composition and environmental factors. hydrobiologia 424: 1–11. naselli-flores l, 2003. man-made lakes in mediterranean semiarid climate: the strange case of dr deep lake and mr shallow lake. hydrobiologia 506-509: 13-21. naselli-flores l, 2010. mediterranean climate and eutrophication of reservoirs: limnological skills to improve management, p. 131-142. in: a. ansari, s. singh gill, g. lanza and w. rast (eds.) eutrophication: causes, consequences and control. springer, dordrecht, the netherlands. naselli-flores l, barone r, 1991. lo zooplancton del lago biviere di cesaro. naturalista siciliano, series iv, 15: 211221. naselli-flores l, barone r, 1994. relationship between trophic state and plankton community structure in 21 sicilian dam reservoirs. hydrobiologia 275:197-205. naselli-flores l, barone r, 1995. phytoplankton and underwater light climate in a hypertrophic reservoir (lake arancio, sicily). giornale botanico italiano 129:1288-1292. naselli-flores l, barone r, 1997. importance of water-level fluctuations on the population dynamics of cladocerans in a hypertrophic reservoir (lake arancio, south-west sicily, italy). hydrobiologia 360:223-232. naselli-flores l, barone r, 1998. phytoplankton dynamics in two reservoirs with different trophic state (lake arancio and lake rosamarina, sicily). hydrobiologia 369:163-178. naselli-flores l, barone r, 2000. phytoplankton dynamics and structure: a comparative analysis in natural and man-made water bodies of different trophic state. hydrobiologia 438:65-74. naselli-flores l, barone r, 2002. limnology of a small, temporary water body: the pond of santa rosalia (sicily, italy). internationale vereinigung für theoretische und angewandte limnologie: verhandlungen 28:1673-1677. naselli-flores l, barone r, 2005. water-level fluctuations in mediterranean reservoirs: setting a dewatering threshold as a management tool to improve water quality. hydrobiologia 548:85-99. naselli-flores l, barone r, 2011. fight on plankton! or, phytoplankton shape and size as adaptive tools to get ahead in the struggle for life. cryptogamie, algologie 32:157-204. naselli-flores l, barone r, 2012. phytoplankton dynamics in permanent and temporary mediterranean waters: is the game hard to play because of hydrological disturbance? hydrobiologia 698:147–159. naselli-flores l, barone r, 2019. mixotrophic phytoplankton dynamics in a shallow mediterranean water body: how to make a virtue out of necessity. hydrobiologia 831:33-41. naselli-flores l, barone r, chorus i, kurmayer r, 2007. toxic cyanobacterial blooms in reservoirs under a semiarid mediterranean climate: the magnification of a problem. environmental toxicology 22:399-404. naselli-flores l, barone r, zunino m, 1998. distribution patterns of freshwater zooplankton in sicilian reservoirs. verhandlungen der internationalen vereinigung fur no nco mm er cia l u se on ly a century of research on micro-organisms from the inland waters of the largest mediterranean island 45 theoretische und angewandte limnologie 26:1973-1980. naselli-flores l, lugliè a, 2014. laghi artificiali dell’italia meridionale e delle isole maggiori. biologia ambientale 28:41-48. naselli-flores l, marrone f, 2019. different invasibility of permanent and temporary water bodies in a semiarid mediterranean island. inland waters 9:411-421. pang w, van de vijver b., 2021. freshwater chrysophycean stomatocysts from monte lauro (buccheri, sicily, italy). phytotaxa 494:177-192. pesce gl, 1988. a new species of schizopera sars 1905 from groundwaters of sicily, italy (crustacea copepoda: diosaccidae). senckenbergiana biologica 68:413-417. pesce gl, 1994. the genus diacyclops kiefer in italy: a taxonomic, ecological and biogeographical up-to-date review (crustacea copepoda cyclopidae). arthropoda selecta 3:13–19. pesce gl, galassi dp, 1987. copepodi di acque sotterranee della sicilia. animalia 14:193-235. pesce gl, galassi dp, apostolov a, 1987. the genus elaphoidella chappuis (copepoda: harpacticoida) in italy, including the description of five new species. italian journal of zoology 54:177-185. pesce gl, galassi dp, cottarelli v, 1988. first representative of the family parastenocarididae from sicily (italy), and description of two new species of parastenocaris kessler (crustacea copepoda: harpacticoida). bulletin zoologisch museum 11:137-142. pieri v, marrone f, martens k, rossetti g, 2020. an updated checklist of recent ostracods (crustacea: ostracoda) from inland waters of sicily and adjacent small islands with notes on their distribution and ecology. the european zoological journal 87:714-740. pieri v, martens k, naselli-flores l, marrone f, rossetti g, 2006. distribution of recent ostracods in inland waters of sicily (southern italy). journal of limnology 65:1-8. platt t, subba rao dv, irwin b, 1985. photosynthesis by picoplancton in the ”lago di faro”, messina (sicily). memorie di biologia marina e di oceanografia 15:61-70. riggio s, 1978. l'ecologia del fiume oreto nel quadro della degradazione ambientale della zona umida di palermo. atti del ii convegno siciliano di ecologia, pp. 175-273. edizioni delphinus, noto, italy. rodríguez lp, granata a, guglielmo l, minutoli r, zagami g, brugnano c, 2013. spring rotifer community structure in the alcantara river (sicily, italy), using different mesh size nets: relation to environmental factors. annales de limnologie – international journal of limnology 49:287-300. roussel jm, paillisson jm, tréguier a, petit e, 2015. the downside of edna as a survey tool in water bodies. journal of applied ecology 52:823-826. ruffo s, stoch f (eds.), 2005. checklist e distribuzione della fauna italiana. 10.000 specie terrestri e delle acque interne. memorie del museo civico di storia naturale di verona, 2^ serie, sezione scienze della vita 16:1-307. russo r, chiaramonte m, palla f, 2001. freshwater cyanobacteria, identified by microscopic and molecular investigations on a colonized fountain surface: a case study in palermo (sicily, italy). conservation science in cultural heritage 21:205-220. saccà a, guglielmo l, bruni v, 2008. vertical and temporal microbial community patterns in a meromictic coastal lake influenced by the straits of messina upwelling system. hydrobiologia 600:89–104. sarà m, 1961. rinvenimento di ephydatia fluviatilis (l.) in sicilia ed osservazioni sugli spongillidi italiani. bollettino di zoologia 28:31-37. schifani e, viviano a, viviano r, naselli-flores l, marrone f, 2019. different lineages of freshwater jellyfishes (cnidaria, olindiidae, craspedacusta) invading europe: another piece of the puzzle from sicily, italy. limnology 20:143-151. sommer u, 1987. factors controlling the seasonal variation in phytoplankton species composition. a case study for a deep, nutrient rich lake, p. 123-178. in: f.e. round and h.h. chapman (eds.), progress in phycological research 5. sorokin jui, donato n, 1975. on the carbon and sulphur metabolism in the meromictic lake faro (sicily). hydrobiologia 47:241-252. taticchi mi, elia ac, rosso a, prearo m, 2011. prime informazioni sulla briozoofauna delle acque interne della sicilia sudorientale. biogeographia 30:327-337. trüper hg, genovese s, 1968. characterization of photosynthetic sulphur bacteria causing red water in lake faro (messina, sicily). limnology and oceanography 13:225-232. vecchioni l, arculeo m, cottarelli v, marrone f, 2021a. rangewide phylogeography and taxonomy of the marine rock pools dweller tigriopus fulvus (fischer, 1860) (copepoda, harpacticoida). journal of zoological systematics and evolutionary research 59:839-857. vecchioni l, arculeo m, marrone f, 2021b. molecular data attest to the occurrence of autochthonous daphnia pulex (crustacea, branchiopoda) populations in sicily. advances in oceanography and limnology 12:1-7. vecchioni l, chirco p, bazan g, marrone f, arizza v, arculeo m, 2021c. first record of temnosewellia minor (platyhelminthes, temnocephalidae) in sicily, with a plea for a reexamination of the identity of the publicly available molecular sequences of the genus. biogeographia 36:a003. vecchioni l, marrone f, naselli-flores l, 2019b. on the occurrence and distribution of calanipeda aquaedulcis kritschagin, 1873 (copepoda, calanoida, pseudodiaptomidae) in sicily, italy, with some notes on coexistence and species replacement in calanoid copepods. advances in oceanography and limnology 10:18-23. vecchioni l, marrone f, rodilla m, belda ej, arculeo m, 2019a. an account on the taxonomy and molecular diversity of a mediterranean rock pool-dwelling harpacticoid copepod, tigriopus fulvus (fischer, 1860) (copepoda, harpacticoida). ciencias marinas 45:59-75. warwick fv, bertola c, 2014. lake physics to ecosystem services: forel and the origins of limnology. limnology and oceanography e-lectures 4:1-47. woolway ri, maberly sc, 2020. climate velocity in inland standing waters. nature climate change 10:1124–1129. zaccone r, caruso g, la ferla r, maimone g, puglisi a, leonardi m, 2000. ciclo annuale dell’attività microbica nell’ecosistema lagunare di tindari (messina). biologia marina mediterranea 7:291-295. no nco mm er cia l u se on ly layout 1 introduction cyanobacteria are the most primitive and ancient photo-oxygenic organisms on earth. with a variety of morphologically different types, they occur in aquatic and terrestrial habitats, as well as in extreme environments such as ultraoligotrophic oceans (seymour, 2014), soil crusts in semi-arid areas (metcalf et al., 2012), and geothermal hot springs (ward and castenholz, 2012). in freshwater ecosystems, they can dominate the phytoplankton assemblages forming blooms in the water column or at the surface of lakes and rivers. this ability to colonise and dominate virtually every habitat on earth is due to their long evolutionary history (tomitani et al., 2006) and to a number of adaptations that render them competitively superior to many eukaryotic phytoplankton groups (dokulil and teubner, 2000). in this regard, cyanobacteria are one of the major eubacterial lineages. the diversity within this group includes morphological characters (from single cells to branching filaments), the formation of specialized cells (akinetes and heterocytes) and cell structures (aerotopes or gas-vesicles), and physiology (nitrogen fixation, heterotrophy, motility) (graham et al., 2009). the growth and dominance of cyanobacteria have important consequences on the characteristics of water bodies. along with effects in common with other eukaryotic algae (decrease of water transparency, increase of ph, alteration of biogeochemical cycles), cyanobacteria have further critical impacts, which include changes in the biomass size spectrum and edibility by zooplankton, and the production of an impressive range of secondary toxic metabolites (sukenik et al., 2015), including hepatotoxins and neurotoxins (metcalf and codd, 2012). the massive development of this algal group is considered as one of the principal problems for the utilisation of water resources, with possible impacts on human health (chorus and bartram, 1999; ibelings et al., 2014). the cyanobacterial species belonging to the nostocales have the ability to develop specialized cells with distinctive morphological and physiological adaptations for advances in oceanography and limnology, 2015; 6(1/2): 33-45 original article doi: 10.4081/aiol.2015.5456 historical colonisation patterns of dolichospermum lemmermannii (cyanobacteria) in a deep lake south of the alps nico salmaso,1* adriano boscaini,1 camilla capelli,1,2 leonardo cerasino,1 manuela milan,1,3 sara putelli,1,4 monica tolotti1 1iasma research and innovation centre, istituto agrario di s. michele all'adige fondazione e. mach, via e. mach 1, 38010 s. michele all’adige (tn), italy; 2department of biology, university of florence, via la pira 4, 50121 firenze, italy; 3department of ecology and environmental science, umeå university, linneus väg 6, 901 87 umeå, sweden; 4department of food, environmental and nutritional sciences, university of milan, via l. mangiagalli 25, 20133 milano, italy *corresponding author: nico.salmaso@fmach.it abstract since the beginning of the 1990s, lake garda showed the appearance of extended surface water blooms of dolichospermum lemmermannii (nostocales, cyanobacteria). between the 1990s and the 2000s, the blooms appeared also in the other large lakes south of the alps (iseo, como and maggiore). despite the sudden appearance of the blooms, the correct identification of the establishment time of the populations of dolichospermum in the southern subalpine lake district remained unclear. in this work, the establishment of the populations of d. lemmermannii in lake garda has been evaluated by the direct counting of sub-fossil akinetes extracted from sediment cores, and by estimating the abundance of filaments germinated from sub-fossil viable akinetes. the two techniques provided comparable results, allowing locating the beginning of the establishment of dolichospermum around the middle of the 1960s. four strains of dolichospermum germinated from akinetes isolated from the core sediments between around the 1989 and 2012 did not show any mutation or recombination signal in the rpob gene sequences, suggesting a strong founder effect. the establishment of dolichospermum coincided with the beginning of the rapid increase of total phosphorus as inferred from the distribution of sub-fossil diatoms in the sediment core. these results supported the hypothesis of a strong link between the shift of lake garda from ultraoligotrophy/oligotrophy to oligo-mesotrophy and the development of dolichospermum. this colonisation pattern was possibly reinforced by the increase in the water temperatures in the subalpine lake district during the last 3 decades. in warmer lakes, gas-vacuolated nostocales are favoured by high replication rates and, in particular, by their ability to control vertical movements in stratified water columns. this allows these species to exploit the gradients of light and nutrients, giving them a competitive advantage compared to other species. from a management point of view, the control and decrease of dolichospermum should be obtained through the reduction and control of nutrient loads to the lake. key words: cyanobacteria; dolichospermum; invasion; lake garda; akinetes; resurrection ecology. received: july 2015. accepted: november 2015. no nco mm er cia l u se on ly 34historical colonisation patterns of dolichospermum in lake garda the fixation of atmospheric nitrogen (heterocytes) and survival in harsh environmental conditions while quiescent in bottom sediments (akinetes). mature akinetes have a normal cell wall surrounded by a thick three layered coat and a conspicuous granulation due to high concentrations of glycogen and cyanophycin, a nitrogen-storage polymer (lee, 2008; meeks et al., 2002). the loss of gas-vesicles and the increase in cytoplasmic density favour the sinking of overwintering akinetes to the bottom sediments. functionally, akinetes act as spores. they differentiate from vegetative cells by increasing in size and thickening the walls, and by drastically reducing photosynthetic and respiratory capabilities. for example, studies carried out on aphanizomenon ovalisporum forti showed that mature akinetes isolated from cultures maintained only residual photosynthetic activity, as indicated by very low values of maximal photosynthetic quantum yields (sukenik et al., 2007). the rearrangement of the photosynthetic apparatus with the maturation of akinetes is essential to enter the dormant period, assuring, at the same time, a quick recovery of photosynthesis after germination. the environmental factors activating the formation of akinetes differ between species and even strains. after the examination of several case studies, kaplan-levy et al. (2010) identified light intensity as the major factor triggering the formation of akinetes, though other factors, such as light quality, temperature, limitation of phosphorus and other nutrients and micronutrients, were considered important in many other instances. on the other hand, the germination of akinetes is controlled by various environmental factors, including temperature, nutrients, day length, penetration of light to the sediments and intensity of turbulence inducing resuspension near the bottom sediments (kaplan-levy et al., 2010). in general, akinete formation takes place at the end of the growth season. reynolds (1972) observed akinetes in water columns after a dense anabaena population had decreased. kim et al. (2005) found a gradual increase in akinete cell densities in the surface sediments after a drop of anabaena flos-aquae brébisson ex bornet & flauhault densities. nevertheless, studying the seasonal dynamics of anabaena flos-aquae, kravchuk et al. (2006) identified two types of akinetes. while in early summer akinetes germinated in water column soon after differentiation, contributing to the vegetative reproduction of populations, in late summer akinetes were deposited to bottom sediments, acting as resting stages. along with other algal remains, the study of akinetes in core sediments may provide more comprehensive knowledge of past lake phytoplankton assemblages and ecological conditions (livingstone, 1984). combined with other paleolimnological proxies, sub-fossil akinetes may contribute to the assessment of the effects of eutrophication and climate change on lake biota. akinetes are very resistant to degradation. fossil akinetes have been identified in 1.6 to 2.1 billion years old cherts in different regions of the world (tomitani et al., 2006). at the beginning of the 1990s, lake garda showed a quick and unexpected appearance of surface blooms of a pelagic nostocales, dolichospermum lemmermannii (p.g. richter) p. wacklin, l. hoffmann & j. komárek. since their first records, the blooms raised major concerns because of the potential toxigenic effects and the impacts on the tourism economy. the exact timing of the colonisation, however, remained uncertain because it was based almost exclusively on the detection of the surface water blooms and on a few scattered phytoplankton analyses available before the 1990s. moreover, since the distribution of dolichospermum is clearly localised in the temperate and boreal regions (salmaso et al., 2015a), the north to south colonisation pattern of this species contrasted with the prevalent south to north invasion routes of tropical species to the european continent (sukenik et al., 2012). on the other hand, many studies demonstrated that the spread of cyanobacteria can be favoured, besides by eutrophication, by the increase of water temperatures caused by global warming (paerl and huisman, 2009; mehnert et al., 2010). in a recent work, (salmaso et al., 2015a) hypothesised that the increase of dolichospermum in lake garda could be ascribed to the increase of nutrients observed during the last 40 years. this hypothesis will be further explored in this study, by comparing the distribution of sub-fossil akinetes of d. lemmermannii in the deep sediments of the lake with the secular trend of phosphorus recently inferred from sub-fossil diatoms by milan et al. (2015). specific objectives of this work include: i) the reconstruction of the historical colonisation patterns of dolichospermum lemmermannii in lake garda based on the analyses of sub-fossil akinetes and its interpretation based on the secular changes in p-availability and longterm changes in water temperatures; ii) the assessment of the comparability of results obtained from the direct counting of sub-fossil akinetes extracted from sediments and the estimation of the abundances of strains germinated from sediments; iii) the phylogenetic confirmation of the taxonomic identity of recent and ancient populations of d. lemmermannii isolated from environmental samples and originated from the germination of akinetes from the sediment layers, respectively. methods study site lake garda is the largest italian lake. it has a volume of 49×109 m3, a surface of 368 km2, and a maximum depth of 350 m. thanks to attractive landscape and water quality, no nco mm er cia l u se on ly 35 n. salmaso et al. the lake is an important resource for recreation and tourism. moreover, its waters are also used for agriculture, industry, fishery and drinking. owing to its climatic location, lake garda should be classified as warm monomictic. in this category, lakes circulate completely once a year in the winter at or above 4°c, showing stable stratification for the remainder of the year. nevertheless, owing to its great depth, complete mixing can occur only irregularly, after cold winters (oligomixis). since the beginning of the 1990s, lake garda has been regularly investigated with monthly samplings in a station located at the deepest point of the basin (45.69 n, 10.72 e). the lake garda sampling station was included in the italian network for the long term ecological research in 2007 (lter, www.lteritalia.it). climatic, environmental, and phytoplankton data mean monthly homogenised air temperatures were obtained from the long-term instrumental climate data recorded in the alpine region in the framework of the project histalp (auer et al., 2007). in this work, we have considered the temperatures recorded at torbole riva del garda, which is the nearest histalp meteorological station to the lake sampling point (ca. 20 km north). since a preliminary interpolation by loess of the long-term temporal development of the air temperatures showed the presence of abrupt changes, the data were analysed by piecewise regressions. the selection of the break-point was guided by computing the residual standard error of several two-segment piecewise models and by the results of the loess interpolation (r core team, 2015). the collection of samples and measurements in the water layers 0-2, 9-10 and 19-21 m were made at monthly intervals since 1991 in the lter station. samples for the analysis of phytoplankton abundance and nutrients were collected since 1993 and 1995, respectively. vertical profiles of water temperature were measured with multi-parameter probes (idronaut ocean seven 401 and 316plus, and seabird sbe 19-03) (sharma et al., 2015). the significance of the linear long-term trend of annual mean water temperatures in the layer 0-50 m was tested with the mann-kendall (mk) test, whereas the trend coefficients were estimated with the non-parametric theil-sen estimate of slope following the procedures described in salmaso and mosello (2010). statistical analyses were carried out with r 3.2.2 (r core team, 2015). nitrate nitrogen (no3-n) and total phosphorus (tp) were analysed using standard methods (cerasino and salmaso, 2012). phytoplankton counting was carried out using the utermöhl method as described in salmaso (2010). sediment sampling, and akinetes analyses a sediment core was collected with a gravity kajak corer (uwitec, mondsee, austria) at the lter station of lake garda on 11 march 2014. the core was vertically extruded and sliced in the laboratory at 0.5 cm intervals from 0 to 30 cm and at 1 cm intervals from 31 cm down to the core bottom (45 cm). hereafter the layers will be indicated by their mid-depth (e.g., 1-1.5 cm: 1.25 cm). water content was determined after drying ~2 g of wet sediment at 105°c for 24 h. the direct microscopic analysis of akinetes is disturbed by the presence of sediments and organic material (including siliceous diatoms). this is especially true in low productivity, nutrient-poor lakes, where the abundance of akinetes is low. in such environments, a reliable counting could require a pre-treatment of the samples aimed at removing carbonate and silicate crystalline material, and diatom carpets. before observation, the sediment samples were therefore processed adapting standard laboratory methods used for pollen analysis (faegri and iverson, 1989). examples of applications of these methods in the analyses of sub-fossil akinetes and cyanobacteria are reported in platt bradbury et al. (2004), danielsen (2009), menozzi et al. (2010) and miras et al. (2015). for each sediment layer, subsamples of ca. 2 g of fresh sediment were put in 15 ml plastic test tubes. the sediment was initially processed with 10% hydrochloric acid (hcl) for the removal of carbonates. for the removal of organic matter, 10 ml 10% koh were added to the test tubes placed in a boiling water bath for 10 min, stirring slowly. the siliceous material was removed by adding hydrofluoric acid (hf). compared with standard methods, the concentrations of hf were reduced to 10% to reduce the risk in laboratory procedures. the fluorosilicates gels formed during the reaction with hf were removed by adding 10 ml hot 10% hcl for 20 min. after the treatment with hcl, koh and hf, the samples were centrifuged at 3000× g for 3-4 minutes, pouring away the supernatant. after the last treatment with hf, the final residue was diluted in 100 ml water and fixed with lugol’s solution. the samples, diluted 1:20, were transferred into10 ml sedimentation chambers and analysed by an inverted microscope at 400×. the enumeration of akinetes was carried out analysing the whole bottom chamber of two replicates (supplementary fig. 1). the application of this method allowed obtaining clean samples, rich in akinetes. furthermore, after this treatment the akinetes were not hidden by siliceous frustules, therefore reducing the risk of misidentification, which is particularly high in lakes rich in diatoms, such as lake garda (salmaso, 2010). during this research, the reliability of the method was tested with a parallel counting of akinetes on treated (t, n=4) and non treated (nt, n=8) fresh sediment samples (without step 2 in supplementary fig. 1) extracted from a core rich in akinetes collected in lake balaton in november 2010 within the framework of the project eulakes (www.eulakes.eu). the concentrations (average no nco mm er cia l u se on ly 36historical colonisation patterns of dolichospermum in lake garda ±se) of akinetes in the t and nt samples were equivalent (83±9 and 88±6 akinetes mg–1 fresh sample, respectively; anova, f1,10=0.23, p>0.60). the germination of filaments from sub-fossil akinetes was obtained using ca. 4 g of fresh sediment inoculated in 100 ml conical flasks. a final volume of 60 ml was obtained by adding asm-1 culture media (carmichael and gorham, 1974). the cultures were added with actidione (cycloheximide, 250 mg l–1) as an inhibitor of eukaryotic growth (livingstone and jaworski, 1980). the flasks were incubated in a growth cabinet at 20°c under continuous irradiance with day light florescent lamps at 85 µmol m–2 s–1. the germination of akinetes was observed up to 16-21 days. later on, the culture media with germinated vegetative cells were harvested and preserved in 100 ml bottles fixed with lugol’s solution. after dilution 1:40, the subsamples were transferred into 10 ml sedimentation chambers. the enumeration of the germinated cells was carried out with an inverted microscope by analysing the whole bottom chamber of 4-6 replicates (supplementary fig. 2). owing to the successive vegetative replication after the initial germination, the final densities of the cells recorded in the flasks were only roughly proportional to the initial number of akinetes in the sediments. on the other hand, one of the advantages of this method is to allow resurrection of ancient akinete-forming populations. core dating and analysis the 2014 sediment core was dated by parallelisation (thompson et al., 2012) with a core collected at the same sampling point in october 2009. the 2014 and the 2009 cores showed a well resolved and comparable macroscopic layering, and very similar depth profiles of water content (r=0.94, p<0.01). the core collected in 2009, and used for the parallelization, was dated analysing radionuclide activity of 210pb, 226ra, 137cs and 241am in subsamples of individual sections, as described in milan et al. (2015). the radionuclides allowed dating the core until the end of 1800s (26-27 cm). the dating of the deeper layers was estimated based on the 14c analysis made on a vegetable remain isolated at 48.5 cm, which was dated back to 1388-1448 ad (milan et al., 2015). in order to assign an age to each sediment layer of the core collected in 2014, a loess interpolation of radiometrically determined dates for non-contiguous sediment layers was performed using r 3.2.2 (r core team, 2015). in this work, we will also report additional results regarding the bacterial photosynthetic pigments and secular lake tp concentrations reconstructed from sub-fossil diatoms obtained from the analyses of the 2009 core (milan et al., 2015). photosynthetic pigments were determined after sonication of sediments, extraction with acetone and analysis by light spectrometry and hplc. sub-fossil diatom frustules were cleaned following standard procedures (battarbee et al., 2001). diatom-inferred total phosphorus concentrations (di-tp) were reconstructed using a weighted-average regression model with inverse deshrinking that was calibrated against the northwest european training set (nweu-tp) (bennion et al., 1996). the detailed description of these procedures was reported in milan et al. (2015). isolation of dolichospermum strains and genetic analyses single filaments of germinated dolichospermum were isolated from the cultures (stage 2 in supplementary fig. 2) obtained from the layers 1.25, 3.25, 4.25 and 6.25, corresponding to the years 2012, 2005, 2000, and 1989, respectively. compared with filaments isolated from environmental samples, the individuals obtained from the germinated akinetes were more difficult to grow. after washing 3-4 times, the single filaments were grown in z8 medium at 20°c under 16:8 h light:dark photoperiod and at 25 µmol m–2 s–1. dna extraction of the cultures containing the single strains was carried out following shams et al. (2014). phylogenetic analyses were carried out by the amplification of the rpob gene encoding the β subunit of rna polymerase. rpob was amplified with the primers rpobanaf and rpobanar, following a modified protocol described in rajaniemi et al. (2005). the reaction mix (20 µl), contained 1x optimized dynazyme pcr buffer, 0.25 mm dntps mix, 0.4 u of dynazyme ii dna polymerase, 0.5 µm of the two primers, and 5-20 ng dna templates. pcr amplification involved an initial denaturation for 5 min at 94°c, 30 cycles of amplification: 1 min at 94°c, 1.5 min at 59°c and 2 min at 68°c, and a final elongation for 7 min at 68°c. pcr products were separated by 1% agarose gel electrophoresis stained with ethidium bromide. the rpob gene was sequenced with the same primers used in the pcr following the methods described in shams et al. (2014). the sequences (564-576 bp long) were deposited to the european nucleotide archive (ena) with accession numbers ln871489ln871492. phylogenetic analysis the rpob sequences of the germinated dolichospermum strains were compared with other 2 sequences of d. lemmermannii isolated from recent pelagic samples collected in lake garda in 2013 and 2014 (accession numbers ln871471 and ln871475, respectively; salmaso et al., 2015a) and other dolichospermum species from genbank. microcystis aeruginosa (kützing) kützing (eu151907) was included as outgroup. all the steps in the phylogenetic analysis were performed with the free statistical software r 3.2.2 (r core team, 2015). sequences were aligned with muscle (edgar, 2004), and no nco mm er cia l u se on ly 37 n. salmaso et al. poorly aligned positions and divergent regions were eliminated using aliscore (misof and misof, 2009). phylogenetic trees were computed by maximum likelihood (ml) with 1000 bootstrap (bs) replicates using the package phangorn. the analysis of dna substitution models indicated in the k80+g the best-fitting evolutionary model for the rpob gene. results long term development of air and water temperatures the long-term development of the mean annual air temperatures recorded at the northern border of the lake showed an overall tendency to increase (fig. 1a). the interpolation by the loess function suggested the existence of at least two different linear relationships in the data, with an abrupt change of slopes around the first half of the 1960s. the piecewise regression (r2=0.60, p<0.001) indicated an increase in air temperature around 0.07°c per decade between the end of the 1800s and 1965, and 0.45°c per decade in the successive years. overall, since 1965, the mean annual air temperatures increased from around 12.5°c to over 14°c. the long term change of the lake water temperatures was evaluated in the mixolimnetic layer, between 0 and 50 m (fig. 1b). since 1991, this layer underwent complete cooling every year, thus representing the part of the lake most affected by the winter climate. in this layer, the annual mean water temperatures showed a significant increase (kendall’s tau 0.34, p<0.05), at a rate of 0.24°c per decade (p<0.05). nutrients in the first half of the 1970s, total phosphorus concentrations measured in the whole water column (0-350 m) were around 5-10 µg p l–1 (fig. 2a). in the following 30 years, tp increased continuously, until reaching concentrations between 20 and 25 µg p l–1 around 2000. more recently (i.e., after 2005), tp showed a slow tendency to decrease. present concentrations are around 18 µg p l–1 (fig. 2a). during the coldest months, in the epilimnetic layers tp reached homogeneous concentrations from the surface to the bottom only during complete overturn. during summer stratification, epilimnetic tp concentrations were generally between 5 and 10 µg p l–1 (salmaso, 2011). the pattern characterizing the long-term temporal changes of tp evaluated on the basis of chemical analyses (fig. 2a) was confirmed by the tp values inferred from the composition of sub-fossil diatoms analysed in the core collected in 2009 (fig. 3). the reconstruction of diatomfig. 1. long-term temporal development of the mean annual values of (a) air temperatures recorded at the histalp meteorological station of torbole riva del garda, around 20 km north the sampling station; the interpolation by the 2-segmented line was obtained by a piece-wise regression. (b) annual mean water temperatures in the layer between 0 and 50 m recorded in the lter station of lake garda; the continuous line shows the linear trend obtained by the theil-sen estimate of slope. in (a) and (b) the series have been smoothed using the loess procedure (dashed lines). no nco mm er cia l u se on ly 38historical colonisation patterns of dolichospermum in lake garda inferred tp concentrations showed that, at least since the last period of the middle ages and until the beginning of the 1960s, lake garda was characterised by near ultraoligotrophic conditions (di-tp <10 µg p l–1). the increase of di-tp began only during the second half of the 1960s, in line with the results obtained by the modern limnological analyses (mosello and giussani, 1997; fig. 2a and fig. 3). though less distinct, this pattern was confirmed also by the tp values inferred by pigment analyses (car-tp; fig. 3). fig. 2. long-term temporal development of the mean annual values of (a) total phosphorus in the whole water column (0350 m), and (b) cyanobacteria and (c) dolichospermum lemmermannii in the trophogenic layers (0-20 m). in (a) the data before 1995 were redrawn from salmaso and mosello (2010). the series have been smoothed using the loess procedure (dashed lines). in (c), the continuous line reports the loess smoothing computed including the outlier. fig. 3. depth profiles of reconstructed phosphorus and sub-fossil cyanobacterial pigments in the sediment core collected from the deepest point of lake garda in 2009. di-tp are tp concentrations reconstructed from sediment diatom abundance data using an inverted weighted averaging procedure based on the north west europe training set (milan et al., 2015). echinenone is a xanthophyll synthesised exclusively by some cyanobacteria. while zeaxanthin is a major pigment in cyanobacteria, it can be present in low quantities in most other eukaryotic phytoplankton species (takaichi, 2011; milan et al., 2015). loi, loss-on-ignition; the box highlights the first establishment of dolichospermum lemmermannii. no nco mm er cia l u se on ly 39 n. salmaso et al. in the whole water column, the average annual concentrations of nitrates (the prevalent n-compound) between 1996 and 2013 ranged between 290 and 400 µg l–1. in the trophogenic layers (0-20 m), annual averages of no3-n generally varied from 200 to 260 µg n l–1, without showing any apparent trend in the analysed period (p»0.1). cyanobacteria since the beginning of phytoplankton analyses, the annual averages of the cyanobacteria biovolumes showed a gradual increase from 50-100 mm3 m–3 until 250-350 mm3 m–3 in the middle of 2000s (fig. 2b). in the successive period, cyanobacteria showed a dramatic decrease, until reaching biovolumes similar to those measured at the beginning of the 1990s. the analyses of phytoplankton allowed inferring the response of cyanobacteria to the changes of phosphorus in the last 20 years. on a multi-decadal scale, zeaxanthin and echinenone, which are carotenoids typically produced by cyanobacteria, showed a sudden increase in the second half of the 1970s (fig. 3). in the previous period, both pigments were measured with very low or undetectable values. the increase of cyanobacteria as inferred from these two proxies was shifted forward by around 10 years compared with the increase of di-tp (fig. 3). it is worth considering that both zeaxanthin and echinenone did show a tendency to decrease during and after the first half of 2000s, therefore confirming the recent tendency of cyanobacterial decrease estimated from the algal counts. dolichospermum lemmermannii: long-term temporal development and water blooms the annual means of dolichospermum lemmermannii biovolume were over one order of magnitude lower than those of total cyanobacteria (fig. 2c). excluding the higher biovolume value computed in 2012, the long-term temporal pattern of d. lemmermannii biovolume was comparable to that of total cyanobacteria. the very low annual average biovolumes of d. lemmermannii at the beginning of the 1990s were consistent with the timing in the appearance of the first visible water bloom. the surface accumulation of filaments of d. lemmermannii was observed for the first time in the shallower and wind sheltered e-basin (salmaso et al., 1994). in the successive years, and quite rapidly, the water blooms expanded over the whole lake, including the northernmost and deeper zone. the blooms, which were observed both in the central zones and along the lake shores, were usually accumulated by the wind in sheltered bays and harbours. the episodes were always observed between early summer and autumn. a careful investigation among the scientists studying the lake since the 1950s and the fisherman allowed confirming the absence of the water blooms of dolichospermum before the 1990s (salmaso et al., 1994). between the second half of the 1990s and 2000s, d. lemmermannii appeared also in the other largest lakes south of the alps, i.e., lakes iseo, como and maggiore (garibaldi et al., 2003; mosello et al., 2010; salmaso, 2010; callieri et al., 2014). historical distribution of akinetes and resurrected populations of dolichospermum the last measurable quantity of sub-fossil akinetes in the core collected in 2014 was found at 16.75 cm (1965; fig. 4a, supplementary fig. 3). repeated counts made on layers corresponding to the previous period, between 1935 and 1961, were negative. since 1965, the number of akinetes remained practically constant until the end of the 1970s. since then, akinetes abundance suddenly and fastly increased until 2007, then began to decrease. the morphology and morphometry of akinetes in the sediments were the same as those found in the pelagic samples, with length and width around 14-20 µm and 6.5-8.5 µm, respectively (salmaso et al., 2015a). the germination of akinetes, with the corresponding growth of resurrected populations of dolichospermum, was observed between the surface and the core depth of 14.25 cm (1971; fig. 4b). similarly to what was observed with the direct counting of akinetes (fig. 4a), the new filaments developed with constant and low abundances until 1985. in the following periods, densities steadily increased till the end of 2000s. the counts corresponding to the uppermost analysed layer (dating 2012) showed a decrease in the cell abundances (fig. 4b). the vegetative cells in the filaments presented characteristics fully compatible with the physiographic characteristics of d. lemmermannii. nevertheless, compared with the natural populations, the formation of akinetes and heterocytes was usually not observed in the sediment cultures. the phylogenetic analyses allowed clearly confirming the taxonomic nature of the sub-fossil populations of dolichospermum germinated from the sediments. the four rpob sequences of the resurrected strains corresponding to the years 1989, 2000, 2005 and 2012 were identical. these four strains formed a compact clade along with the strains of d. lemmermannii isolated in the pelagic samples of lake garda and other two strains isolated in n-europe (fig. 5). the other dolichospermum species formed different separated clades. discussion the analyses of the distribution of the sub-fossil akinetes preserved in a sediment core collected in 2014 allowed to reconstruct the colonisation pattern of dolichospermum lemmermannii in lake garda. the clear no nco mm er cia l u se on ly 40historical colonisation patterns of dolichospermum in lake garda identification of the beginning of the colonisation is a necessary requirement to interpret the causes that favoured the introduction and the successive rapid spread of this nostocales in lake garda and in the whole southern subalpine lake district. the two laboratory procedures used to estimate the quantitative presence of dolichospermum in the sediment layers provided consistent and reproducible results. the direct determination of akinetes in cleaned sediments and the counting of the cells germinated from viable akinetes indicated the beginning of the colonisation in the middle of the 1960s and in 1971, respectively. though the discrepancy between the two methods, around 5 years, was very limited we cannot exclude the possible presence of a few other akinetes also in the older sediment layers. nevertheless, both the consistency of the two methods in detecting the presence of akinetes, and the very rapid increase of akinetes after the 1970s, strongly suggest that the populations of dolichospermum increased their importance during the 1980s and 1990s (fig. 4), i.e., during the period coinciding with the appearance of the first surface blooms (salmaso et al., 1994). adopting the terminology by williamson and fitter (1996), the methods used in this work allowed to identify the period of establishment of self-sustaining and naturalized populations, after a first importing and introduction of the species. the analysis of the records of sub-fossil akinetes preserved in deep layers of lake sediments represents a powerful and efficient tool to reconstruct the long-term temporal development of potentially toxic nostocales in water bodies at secular timescales. akinetes can be prefig. 4. vertical distribution of (a) sub-fossil akinetes and (b) cells developed and reproduced from the germination of viable akinetes of dolichospermum lemmermannii preserved in the sediment layers. the densities have been transformed by square root. the horizontal dashed lines indicate the period of first appearance of the blooms. in (b), the isolation and resurrection in culture conditions of single ancient dolichospermum strains was obtained from the layers indicated with the arrows, which correspond to the years 2012, 2005, 2000, and 1989. the counting of akinetes in the sediments and the counting of the cells originated from the germination of akinetes indicated the beginning of the colonisation of dolichospermum in the middle of the 1960s and in 1971, respectively. fig. 5. maximum likelihood (ml) rooted topology of the four resurrected dolichospermum lemmermannii strains isolated from the sediments of lake garda (1989-2012) and other cyanobacteria (identified by names and accession numbers) based on alignment of the rpob gene. the d. lemmermannii strains isolated from the pelagic samples collected in 2013 and 2014 in lake garda have accession numbers ln871471 and ln871475, respectively. bootstrap values <70 were not shown. no nco mm er cia l u se on ly 41 n. salmaso et al. served for very long times in the sediments. van geel et al. (1994) found that, since 1000 ad on, sub-fossil akinetes of aphanizomenon and anabaena were present in enormous quantities in the deep sediment layers of lake gosciaz, in poland. the increase of nostocales was interpreted as the effect of the intensification of farming and land fertilization, which triggered and intensified the eutrophication of the lake. similarly, peaks of akinetes of anabaena (dolichospermum) in the ancient sediments of lake aydat, were found since the 5th century ad (miras et al., 2015). in lake kirmanjärvi, finland, kauppila et al. (2012) utilised the mutual changes in the abundance of the akinetes produced by anabaena and aphanizomenon to infer the long-term trophic status changes since the mid 16th century. akinetes are markedly more resistant than vegetative cells to a wide range of extreme physical, chemical and biotic disturbances (fay, 1988). nevertheless, the resistance against specific stressors can be different between species. for example, while the akinetes of anabaena cylindrica lemmermann were resistant to desiccation, hot temperatures and sunlight (hori et al., 2003), the akinetes of anabaena circinalis rabenhorst ex bornet & flahault were susceptible as vegetative cells to desiccation or ultraviolet radiation (fay, 1988). when buried in the sediments, the akinetes can retain their viability for decades. in this study, we use the term viability as capability to germinate and to produce new filaments and populations once given the appropriate conditions. in lake okaro, germination experiments demonstrated that akinetes of aphanizomenon issatschenkoi (usačev) proshkina-lavrenko retained their ability to germinate in core sediments dating 120 bp (wood et al., 2009). livingstone and jaworski (1980) demonstrated that akinetes of aphanizomenon flos-aquae ralfs ex bornet & flahault and dolichospermum (anabaena) lemmermannii isolated from 18 and 64 years old sediments were able to germinate, giving rise to viable populations. in the case of aphan. flos-aquae, the inability to germinate was observed in spite of a large abundance of akinetes in layers older than 18 years. on the one hand, that study demonstrated that akinetes not only have overwintering functions, but are capable of contributing to the long-term survival of nostocales in lakes. on the other hand, it showed a different ability of akinetes produced by different species to germinate giving rise to mature populations. the implications are clear, suggesting that the germination of akinetes cannot be used to compare the long-term changes in different species of nostocales. in lake garda, the concordance in the temporal distribution of sub-fossil akinetes and germinated cells indicates that the akinetes of d. lemmermannii retain the ability to germinate in sediments of at least 35 years old. to date, it is not possible to evaluate how much this approximation is underestimated, because the deepest layer of germinability (fig. 4b) roughly coincides with the limit of akinete presence (fig. 4a). since akinetes can be resurrected, the genetic structure of sub-fossil populations can be compared with that of modern strains, allowing a more robust assessment of historical fluctuation patterns. in principle, the use of methods of resurrection ecology might allow to reconstruct evolution of organisms by comparing the strains originated from resistance forms with their present descendants (kerfoot and lawrence, 2004). the comparison of strains along temporal gradients could help reveal whether direct effects of climate change or other anthropogenic stressors have caused micro-evolutionary processes (angeler, 2007). the results obtained from sequencing the rpob gene in populations of dolichospermum revitalized in the lake sediments spanning 23 years (1989-2012) and isolated from recent pelagic samples did not show any mutation or recombination signal. actually, previous analyses of other cyanobacteria species (planktothrix rubescens (de candolle ex gomont) k. anagnostidis & j. komárek) in the large lakes south of the alps proved that local clonal expansion and recombination processes were influential in affecting among-lake diversity (d’alelio et al., 2013). present investigations made on the 16s rrna and rpob genes of several d. lemmermannii isolated from environmental samples are showing the existence of different local clonal populations living in the perialpine lakes (salmaso et al., 2015a) and in europe (capelli et al., unpublished data). overall, though based on a very limited number of strains, these results could allow to hypothesise a strong founder effect following the beginning of the establishment of d. lemmermannii (between the half of the 1960s and the end of 1970s) and the successive development of the species in lake garda. possibly, these characteristics could allow identifying the geographical path of colonisation by comparing the local clones with those isolated in other european water bodies. it is widely recognised that the environmental factors mostly affecting the development of cyanobacteria in lakes are eutrophication and high water temperatures (istvánovics, 2009; o’neil et al., 2012; paerl and paul, 2012). the increase of cyanobacteria following nutrient enrichment (p and n) has been described in a wide number of studies (see dokulil and teubner, 2000), and confirmed by experimental works (schindler, 2012). the response is strongly dependent from the autecology of the single species, and the distinction is particularly apparent when considering the ability to fix nitrogen (levine and schindler, 2011). nevertheless, more than eutrophication, a number of recent studies highlighted the role of global warming in favouring the recent spread of nostocales (e.g., cylindrospermopsis and aphanizomenon) from tropical to subtropical and temperate regions (sukenik et al., 2012). no nco mm er cia l u se on ly 42historical colonisation patterns of dolichospermum in lake garda in warmer environments this group, along with chroococcales, is favoured by its high replication rates and by the ability to control the vertical positioning in well stratified water columns induced by light availability and nutritional conditions (e.g., carbohydrates content) (oliver and walsby, 1984; reynolds, 2006). the control of the vertical position allows exploiting the full gradients of light and nutrients, giving to nostocales and other gas-vacuolated species a chief competitive advantage compared to eukaryotic taxa. with the exception of a few localized regions, lake warming was clearly observed and documented all over the world (sharma et al., 2015). besides lake garda (fig. 1b), the increase in water temperatures was confirmed also in other lakes of the southern perialpine area. long-term measurements carried out since the 1970s at spring overturn in lakes garda, iseo, como and maggiore demonstrated an increase in water temperatures ranging from 0.11 to 0.21°c decade–1 (salmaso and mosello, 2010). these data are confirmed also by models simulating the effects of a climate change scenario in lake iseo (valerio et al., 2015). along with the rapid increase of air temperatures recorded during the 1960s, these data could suggest an important role played by climate warming in the establishment and growth of dolichospermum in lake garda. nevertheless, compared to many other nostocales (sukenik et al., 2015), the geographical area of d. lemmermannii was shown to be circumscribed to the temperate and northernmost countries (komárková, 1988), whereas this species was never found in tropical regions (komárek, 2013). conversely, the establishment of this nostocales in the 1960s does appear rather linked to the increase of phosphorus and to the shift of the lake from a state between ultra-oligotrophy and oligotrophy to oligomesotrophy. the ecological requirements of d. lemmermannii have been well described (cronberg, 1999; padisák et al., 2008; reynolds et al., 2002; salmaso et al., 2015b), conforming completely to this assumption. in fact, this species is well adapted to clear, deep stratifying oligomesotrophic lakes, or mesotrophic, non turbid shallow lakes. therefore, we could hypothesise that the establishment of this species in lake garda might be favoured primarily by the increase of anthropogenic pressure and changes in trophic status. the successive development of this species could have been reinforced by the climate warming in the subalpine area and by the consequent increase in water temperatures. the absence of any consistent change in the long term-availability of no3-n since the 1990s suggests a possible minor role played by nitrogen. nevertheless, as highlighted in previous works (salmaso, 2010), owing to the n-fixing abilities of heterocytous species, the low summer concentrations of n that are usually found in the upper surface layers in lake garda could be a positive factor for the growth of dolichospermum in the warmer months. as demonstrated by the analysis of the long-term phytoplankton data (salmaso et al., 2015a), the biovolume values of d. lemmermannii in the trophogenic layers (020 m) of lake garda were always low (<50 mm3 m–3), contributing for a very small fraction to the total annual phytoplankton and cyanobacterial biovolumes (less than 1% and 5%, respectively; see also fig. 2 b,c). the dominant cyanobacterial species in lake garda and in the other large lakes south of the alps is in fact planktothrix rubescens (salmaso et al., 2012; leoni et al., 2014). on the other hand, since the blooms are caused by a rapid vertical movement of filaments present in the euphotic layers, these accumulations provide a false idea of massive cyanobacterial development and growth in the surface waters. these considerations confirm that d. lemmermannii should be considered a well established species rather than an invasive or pest species (williamson and fitter, 1996; keller et al., 2011). we could anticipate that, if the downward trend of p in lake garda will be confirmed, then we could expect a further decrease of cyanobacteria and nostocales, and a decrease in the frequency of summer water blooms of dolichospermum. overall, this should favour the recovery of the lake to near-pristine conditions. conversely, a further increase in the trophic status, especially if associated with an additional increase in summer water temperatures, could increase cyanobacteria abundances and surface water blooms. however, it is worth to highlight that an excessive increase in nutrients is a deterrent for the development of d. lemmermannii, which would be replaced by other cyanobacteria better adapted to mesotrophic and eutrophic conditions, such as different species of dolichospermum, as well as aphanizomenon, microcystis and, provided a sufficient amount of light is present in the metalimnion, p. rubescens. these considerations could be confirmed by the absence of consistent populations and blooms of d. lemmermannii in lake lugano. the reasons are due to the lower water temperatures and higher mixing that distinguish the epilimnion of this lake (salmaso et al., 2012) and to the higher trophic status of the trophogenic layers. in fact, the dominant nostocales in lake lugano is aphan. flos-aquae, a species that is not only well adapted to develop under eutrophic conditions (reynolds et al., 2002; salmaso et al., 2015b), but that it is also well fitted to grow in mixed surface layers (kangro et al., 2007). conclusions in this work, we investigated the timing of establishment of the populations of d. lemmermannii in lake garda by the direct counting of sub-fossil akinetes and by the estimation of the abundances of strains germinated from viable akinetes conserved in the sediments. the two no nco mm er cia l u se on ly 43 n. salmaso et al. techniques provided comparable results, allowing locating the beginning of the establishment of dolichospermum between the middle of the 1960s and the early 1970s, respectively. the sequencing of the rpob genes in populations of dolichospermum revitalized in the lake sediments between around 1989 and 2012 and isolated from recent pelagic samples did not show any mutation or recombination, suggesting a strong founder effect following the successive development of the species. the establishment of dolichospermum in lake garda has been linked to the rapid increase of tp in the lake since the 1960s, as inferred from the study of the distribution of sub-fossil diatoms in the core sediment layers. the significant increase in the air and water temperatures observed in the last decades did reinforce and probably increased the effects triggered by anthropogenic eutrophication. if the downward trend of phosphorus that was observed in lake garda since the half of the 2000s will continue, then the biomass of cyanobacteria and nostocales and the frequency of summer water blooms of dolichospermum should show a further and consistent decrease. acknowledgments investigations in lake garda and in the other large lakes south of the alps were made in the framework of the lter (long term ecological research) italian network, site “southern alpine lakes”, it08-000-a (http://www.lteritalia.it/). the activity was supported by phd fellowships to c.c. from the e. mach foundation istituto agrario di s. michele all’adige. we thank the european cooperation in science and technology cost action es1105 cyanocost for networking and knowledge-transfer support. the core sediment in lake balaton was collected in the framework of the eu central europe project eulakes (2ce243p3). we are grateful to our colleagues in fem for their support in the field and laboratory activities, and to the arpa veneto for logistic support in the field. references angeler dg, 2007. resurrection ecology and global climate change research in freshwater ecosystems. j. n. am. benthol. soc. 26:12-22. auer i, böhm r, jurkovic a, lipa w, orlik a, potzmann r, schöner w, ungersböck m, matulla c, briffa k, jones p, efthymiadis d, brunetti m, nanni t, maugeri m, mercalli l, mestre o, moisselin j-m, begert m, müller-westermeier g, kveton v, bochnicek o, stastny p, lapin m, szalai s, szentimrey t, cegnar t. dolinar m, gajic-capka m, zaninovic k, majstorovic z, nieplova e, 2007. histalp historical instrumental climatological surface time series of the greater alpine region. int. j. climatol. 27:17-46. battarbee r, jones v, flower r, cameron n, bennion h, carvalho l, juggins s, 2001. diatoms, p. 155-202. in: j. smol, h.j. birks, w. last, r. bradley and k. alverson (eds.), tracking environmental change using lake sediments. 3. springer bennion h, juggins s, anderson nj, 1996. predicting epilimnetic phosphorus concentrations using an improved diatom-based transfer function and its application to lake eutrophication management. environ. sci. technol. 30:2004-2007. callieri c, bertoni r, contesini m, bertoni f, 2014. lake level fluctuations boost toxic cyanobacterial “oligotrophic blooms”. plos one 9:e109526. carmichael ww, gorham pr, 1974. an improved method for obtaining axenic clones of planktonic blue-green algae 2. j. phycol. 10:238-240. cerasino l, salmaso n, 2012. diversity and distribution of cyanobacterial toxins in the italian subalpine lacustrine district. oceanol. hydrobiol. stud. 41:54-63. chorus i, bartram j, 1999. toxic cyanobacteria in water: a guide to their public health consequences, monitoring and management. available from: http://www.who.int/water_sanitation_health/resourcesquality/toxcyanbegin.pdf cronberg g, 1999. qualitative and quantitative investigations of phytoplankton in lake ringsjon, scania, sweden. hydrobiologia 404:27-40. d’alelio d, salmaso n, gandolfi a, 2013. frequent recombination shapes the epidemic population structure of planktothrix (cyanoprokaryota) in italian subalpine lakes. j. phycol. 49:1107-1117. danielsen r, 2009. dissimilarities in the recent histories of two lakes in portugal explained by local-scale environmental processes. j. paleolimnol. 43:513-534. dokulil mt, teubner k, 2000. cyanobacterial dominance in lakes. hydrobiologia 438:1-12. edgar rc, 2004. muscle: multiple sequence alignment with high accuracy and high throughput. nucleic acids res. 32:1792-7. faegri k, iverson j, 1989. textbook of pollen analysis. blackburn press, caldwell: 340 pp. fay p, 1988. viability of akinetes of the planktonic cyanobacterium anabaena circinalis. p. roy. soc. lond. b bio. 234:283-301. graham le, graham mj, wilcox lw, 2009. algae, 2nd ed. benjamin cummings, san francisco: 720 pp. hori k, okamoto j, tanji y, unno h, 2003. formation, sedimentation and germination properties of anabaena akinetes. biochem. eng. j. 14:67-73. ibelings bw, backer lc, kardinaal wea, chorus i, 2014. current approaches to cyanotoxin risk assessment and risk management around the globe. harmful algae 40:63-74. istvánovics, v, 2009. eutrophication of lakes and reservoirs, p. 157-164. in: g.e. likens (ed.), encyclopedia of inland waters. elsevier. kangro k, olli k, tamminen t, lignell r, 2007. species-specific responses of a cyanobacteria-dominated phytoplankton community to artificial nutrient limitation in the baltic sea. mar. ecol.-prog. ser. 336:15-27. kaplan-levy rn, hadas o, summers ml, rücker j, sukenik a, 2010. akinetes: dormant cells of cyanobacteria p. 5-27. in: e. lubzens, j. cerdà and m.s. clark (eds.), dormancy and resistance in harsh environments. springer. no nco mm er cia l u se on ly 44historical colonisation patterns of dolichospermum in lake garda kauppila t, kanninen a, viitasalo m, räsänen j, meissner k, mattila j, 2012. comparing long term sediment records to current biological quality element data implications for bioassessment and management of a eutrophic lake. limnologica 42:19-30. keller rp, geist j, jeschke jm, kühn i, 2011. invasive species in europe: ecology, status, and policy. environ. sci. eur. 23:23. kerfoot wc, lawrence jw, 2004. experimental paleoecology (resurrection ecology): chasing van valen’s red queen hypothesis. limnol. oceanogr. 49, 1300-1316. kim bh, lee ws, kim y-o, lee h-o, han m-s, 2005. relationship between akinete germination and vegetative population of anabaena flos-aquae (nostocales, cyanobacteria) in seokchon reservoir (seoul, korea). arch. für hydrobiol. 163, 49-64. doi:10.1127/0003-9136/2005/0163-0049 komárek j, 2013. cyanoprokariota. part 3: heterocytous genera, in: süßwasserflora von mitteleuropa, band 19/3. springer. komárková j, 1988. morphological variation in natural populations of anabaena lemmermannii in respect to existence of anabaena utermoehlii. algol. stud. hydrobiol. suppl. vol. 50/53:93-108. kravchuk es, ivanova ea, gladyshev mi, 2006. seasonal dynamics of akinetes of anabaena flos-aquae in bottom sediments and water column of small siberian reservoir. aquat. ecol. 40:325-336. lee re, 2008. phycology, 4th ed. cambridge university press, cambridge: 560 pp. leoni b, garibaldi l, gulati r, 2014. how does interannual trophic variability caused by vertical water mixing affect reproduction and population density of the daphnia longispina group in lake iseo, a deep stratified lake in italy? inland waters 4:193-203. levine sn, schindler dw, 2011. influence of nitrogen to phosphorus supply ratios and physicochemical conditions on cyanobacteria and phytoplankton species composition in the experimental lakes area, canada. can. j. fish. aquat. sci. 56:451-466. livingstone d, 1984. the preservation of algal remains in recent lake sediments, p. 191-20. in: e.y. haworth and j.w.g. lund (eds.), lake sediments and environmental history. leicester university press. livingstone d, jaworski ghm, 1980. the viability of akinetes of blue-green algae recovered from the sediments of rostherne mere. brit. phycol. j. 15:357-364. meeks jc, campbell el, summers ml, wong fc, 2002. cellular differentiation in the cyanobacterium nostoc punctiforme. arch. microbiol. 178:395-403. mehnert g, leunert f, cirés s, johnk kd, rucker j, nixdorf b, wiedner c, 2010. competitiveness of invasive and native cyanobacteria from temperate freshwaters under various light and temperature conditions. j. plankton res. 32:1009-1021. menozzi bi, zotti m, montanari c, 2010. a non-pollen palynomorphs contribution to the local environmental history in the ligurian apennines: a preliminary study. veg. hist. archaeobot. 19:503-512. metcalf js, codd ga, 2012. cyanotoxins, p. 651-675. in: b.a. whitton (ed.), ecology of cyanobacteria ii. their diversity in space and time. springer. metcalf js, richer r, cox pa, codd ga, 2012. cyanotoxins in desert environments may present a risk to human health. sci. total environ. 421/422:118-123. milan m, bigler c, salmaso n, guella g, tolotti m (2015). multiproxy reconstruction of a large and deep subalpine lake’s ecological history since the middle ages. j. great lakes res. (in press). miras y, beauger a, lavrieux m, berthon v, serieyssol k, andrieu-ponel v, ledger pm, 2015. tracking long-term human impacts on landscape, vegetal biodiversity and water quality in the lake aydat catchment (auvergne, france) using pollen, non-pollen palynomorphs and diatom assemblages. palaeogeogr. palaeoclimatol. palaeoecol. 424: 76-90. misof b, misof k, 2009. a monte carlo approach successfully identifies randomness in multiple sequence alignments: a more objective means of data exclusion. syst. biol. 58:21-34. mosello r, giussani g, 1997. [evoluzione recente della qualità delle acque dei laghi profondi subalpini].[book in italian]. doc. ist. ital. idrobiol. 61:1-228. mosello r, ambrosetti w, arisci s, bettinetti r, buzzi f, calderoni a, carrara e, de bernardi r, galassi s, garibaldi l, leoni b, manca m, marchetto a, morabito g, oggioni a, pagnotta r, ricci d, rogora m, salmaso n, simona m, tartari g, veronesi m, volta p, 2010. [evoluzione recente della qualità delle acque dei laghi profondi subalpini (maggiore, lugano, como, iseo e garda) in risposta alle pressioni antropiche e alle variazioni climatiche].[article in italian]. biol. ambient. 24:167-177. o’neil jm, davis tw, burford ma, gobler cj, 2012. the rise of harmful cyanobacteria blooms: the potential roles of eutrophication and climate change. harmful algae 14:313-334. oliver rl, walsby ae, 1984. direct evidence for the role of light-mediated gas vesicle collapse in the buoyancy regulation of anabaena flos-aquae (cyanobacteria). limnol. oceanogr. 29:879-886. padisák j, crossetti lo, naselli-flores l, 2008. use and misuse in the application of the phytoplankton functional classification: a critical review with updates. hydrobiologia 621:1-19. paerl hw, huisman j, 2009. climate change: a catalyst for global expansion of harmful cyanobacterial blooms. environ. microbiol. rep. 1:27-37. paerl hw, paul vj, 2012. climate change: links to global expansion of harmful cyanobacteria. water res. 46:1349-63. platt bradbury j, colman sm, reynolds rl, 2004. the history of recent limnological changes and human impact on upper klamath lake, oregon. j. paleolimnol. 31:151-165. r core team, 2015. r: a language and environment for statistical computing. r foundation for statistical computing, vienna, austria. rajaniemi p, hrouzek p, kastovská k, willame r, rantala a, hoffmann l, komárek j, sivonen k, 2005. phylogenetic and morphological evaluation of the genera anabaena, aphanizomenon, trichormus and nostoc (nostocales, cyanobacteria). int. j. syst. evol. microbiol. 55:11-26. reynolds cs, 1972. growth, gas vacuolation and buoyancy in a natural population of a planktonic blue-green alga. freshwater biol. 2:87-106. reynolds cs, 2006. the ecology of phytoplankton. cambridge university press, cambridge: 552 pp. reynolds cs, huszar v, kruk c, naselli-flores l, melo s, 2002. towards a functional classification of the freshwater phytoplankton. j. plankton res. 24:417-428. no nco mm er cia l u se on ly 45 n. salmaso et al. salmaso n, 2010. long-term phytoplankton community changes in a deep subalpine lake: responses to nutrient availability and climatic fluctuations. freshwater biol. 55:825-846. salmaso n, 2011. interactions between nutrient availability and climatic fluctuations as determinants of the long-term phytoplankton community changes in lake garda, northern italy. hydrobiologia 660:59-68. salmaso n, buzzi f, garibaldi l, morabito g, simona m, 2012. effects of nutrient availability and temperature on phytoplankton development: a case study from large lakes south of the alps. aquat. sci. 74:555-570. salmaso n, capelli c, shams s, cerasino l, 2015a. expansion of bloom-forming dolichospermum lemmermannii (nostocales, cyanobacteria) to the deep lakes south of the alps: colonization patterns, driving forces and implications for water use. harmful algae 50:76-87. salmaso n, cavolo f, cordella p, 1994. fioriture di anabaena e microcystis nel lago di garda. eventi rilevati e caratterizzazione dei periodi di sviluppo. acqua aria 1:17-28. salmaso n, mosello r, 2010. limnological research in the deep southern subalpine lakes: synthesis, directions and perspectives. adv. oceanogr. limnol. 1:29-66. salmaso n, naselli-flores l, padisák j, 2015b. functional classifications and their application in phytoplankton ecology. freshwater biol. 60:603-6019. schindler dw, 2012. the dilemma of controlling cultural eutrophication of lakes. proc. biol. sci. 279:4322-33. seymour jr, 2014. a sea of microbes: the diversity and activity of marine microorganisms. microbiol. aust. 35:183-187. shams s, capelli c, cerasino l, ballot a, dietrich dr, sivonen k, salmaso n, 2014. anatoxin-a producing tychonema (cyanobacteria) in european waterbodies. water res. 69: 68-79. sharma s, gray dk, read js, o’reilly cm, schneider p, qudrat a, gries c, stefanoff s, hampton se, hook s, lenters jd, livingstone dm, mcintyre pb, adrian r, allan mg, anneville o, arvola l, austin j, bailey j, baron js, brookes j, chen y, daly r, dokulil m, dong b, ewing k, de eyto e, hamilton d, havens k, haydon s, hetzenauer h, heneberry j, hetherington al, higgins sn, hixson e, izmest’eva lr, jones bm, kangur k, kasprzak p, köster o, kraemer bm, kumagai m, kuusisto e, leshkevich g, may l, macintyre s, müller-navarra d, naumenko m, noges p, noges t, niederhauser p, north rp, paterson am, plisnier p-d, rigosi a, rimmer a, rogora m, rudstam l, rusak ja, salmaso n, samal nr, schindler de, schladow g, schmidt sr, schultz t, silow ea, straile d, teubner k, verburg p, voutilainen a, watkinson a, weyhenmeyer ga, williamson ce, woo kh, 2015. a global database of lake surface temperatures collected by in situ and satellite methods from 1985-2009. sci. data 2:150008. sukenik a, beardall j, hadas o, 2007. photosynthetic characterization of developing and mature akinetes of aphanizomenon ovalisporum (cyanoprokaryota). j. phycol. 43:780-788. sukenik a, hadas o, kaplan a, quesada a, 2012. invasion of nostocales (cyanobacteria) to subtropical and temperate freshwater lakes physiological, regional, and global driving forces. front. microbiol. 3:86. sukenik a, quesada a, salmaso n, 2015. global expansion of toxic and non-toxic cyanobacteria: effect on ecosystem functioning. biodivers. conserv. 24:889-908. takaichi s, 2011. carotenoids in algae: distributions, biosyntheses and functions. mar. drugs 9:1101-1118. thompson r, clarck m, boulton s, 2012. core correlation, p. 415-430. in: h.j. birks, a.f. lotter, s. juggins and j.p. smol (eds.), tracking environmental change using lake sediments. 5. springer. tomitani a, knoll ah, cavanaugh cm, ohno t, 2006. the evolutionary diversification of cyanobacteria: molecular-phylogenetic and paleontological perspectives. p. natl. acad. sci. usa 103:5442-7. valerio g, pilotti m, barontini s, leoni b, 2015. sensitivity of the multiannual thermal dynamics of a deep pre-alpine lake to climatic change. hydrol. process. 29:767-779. van geel b, mur lr, ralska-jasiewiczowa m, goslar t, 1994. fossil akinetes of aphanizomenon and anabaena as indicators for medieval phosphate-eutrophication of lake gosciaz (central poland). rev. palaeobot. palynol. 83:97-105. ward dm, castenholz rw, 2012. cyanobacteria in geothermal habitats, p. 39-63. in: b.a. whitton (ed.), ecology of cyanobacteria ii. their diversity in space and time. springer. williamson m, fitter a, 1996. the varying success of invaders. ecology 77:1661-1666. wood sa, jentzsch k, rueckert a, hamilton dp, cary sc, 2009. hindcasting cyanobacterial communities in lake okaro with germination experiments and genetic analyses. fems microbiol. ecol. 67:252-60. no nco mm er cia l u se on ly layout 1 introduction the growth of human population and the consequent increasing anthropogenic impact on terrestrial and aquatic ecosystems are today societal problems perceived globally and no longer of interest for the sole scientific community. indeed, the intensive use of the hydrosphere and of non-renewable resources, the globalization of markets and maritime transport, and other socioecological issues, are changing the structure and functions of aquatic ecosystems and their ability to produce goods and services that are useful to humans. inland waters contribute to the supply of clean water by absorbing or filtering pollutants such as heavy metals, excess nutrients, and pesticides and storing water that could otherwise lead to floods (postel and carpenter, 1997). these ecosystems are vulnerable to climate change, land-use reclamation, pollution, salinization and, ultimately, biological invasions (williams, 2001; likens, 2009; vilà et al., 2010). the coastal and deep marine environments are major providers of goods and services (https://ec.europa.eu/maritimeaffairs/policy/blue_growth_ en), such as food, fossil fuels, construction material, transport and recreation (pauli, 2010; silver et al., 2015). they also host invaluable biodiversity that shapes complex ecosystems and play a fundamental role in planetary biogeochemical cycles (e.g. boero et al., 2007; howarth et al., 2011; gamfeldt et al., 2015; d’alelio et al., 2016). altogether, marine ecosystems are exposed to a plethora of threats (crain et al 2008). collectively, all kinds of aquatic systems are threatened by demographic and economic growth via multiple human activities, such as agriculture, fisheries, aquaculture, shipping, urbanization and tourism (dornburg et al., 2008; bogdan et al., 2014). water issues represent an increasing societal challenge, since the 70% of planet earth is covered by water, and only less than 3% of this amount is freshwater. nevertheless, the two thirds of the 35×106 km3 of freshwater resources of the planet are trapped at the ice caps and glaciers, whereas only one third is present in advances in oceanography and limnology, 2018; 9(1): 19-35 article doi: 10.4081/aiol.2018.7580 this work is licensed under a creative commons attribution-noncommercial 4.0 international license (cc by-nc 4.0). present and future of aquatic sciences: the perspective of aiol scientific community for a priority roadmap over the next five years francesca alvisi,1* domenico d’alelio2 1institute of marine sciences, national research council, section of bologna, via p. gobetti 101, bologna; 2department of integrative marine ecology, stazione zoologica anton dohrn. villa comunale, 80121 naples, italy *corresponding author: f.alvisi@ismar.cnr.it abstract water issues represent an increasing societal challenge, since the 70% of planet earth is covered by water. the intensive use of the hydrosphere is changing the structure and functions of aquatic ecosystems and their ability to produce goods and services that are useful to humans. it is therefore necessary that the scientific community makes citizens aware of the results of scientific research on these issues and informs them about the need to intensify the study of the mechanisms that underlie the ongoing changes in aquatic ecosystems. to help meet this need, within the 23rd congress of the italian association of oceanology and limnology (aiol; http://www.aiol.info/), entitled “functioning, alteration and recovery of aquatic ecosystems: the aquatic sciences to understand global change and to make the citizens aware of it” (cagliari, italy, 26-29 september 2017), all participants, among which some renown experts in the field of aquatic sciences, were invited to give their contribution, via a shared and bottom-up built questionnaire, in assessing a set of actions needed to achieve an adaptive and proactive management of changes that the aquatic sciences are going to face in the next five years. the results of this survey allowed us to identify a set of priorities that funding agencies should include in their economic and financial planning in the next future. among all, we pinpoint that there is an urgent need in: (i) promoting sustainable food production by exploiting aquatic systems; (ii) diffusing an opportune spatial planning integrating ecosystem-based management approaches; (iii) developing recovery/remediation plans for contaminated sites; iv) promoting conservation of ecosystems by assessing their conservation status, first of all the water/ecosystem quality; (v) fostering the technological development of sustainable and integrated tools and procedures for environmental monitoring; (vi) developing a better forecasting capacity, particularly of extreme events, by implementing long-term research networks; and, ultimately, (vii) supporting a wider society learning processes and a more effective transfer of knowledge from science to society. key words: aquatic sciences, roadmap, marine sciences, limnology, oceanography, aiol. received: may 2018. accepted: june 2018. no nco mm er cia l u se on ly f. alvisi and d. d’alelio20 liquid form. the largest part of this vital resource is stored in the rocky slopes and underground water basins, with the result that only 0.2 ×106 km3 is available at the earth’s surface (pearce, 2006). in the european union, water scarcity and droughts already affect one third of the territory, while 44% out of the total abstraction of freshwater is used to cool thermal power plants, and 24% for irrigation. as water scarcity and droughts regularly affect large parts of the european territory, water availability and its efficient use are also issues that need to be addressed in europe (alpine convention, 2009; https://ec.europa.eu/jrc/en/research-topic/water). in addition, chemical compounds used in our daily activities can travel along rivers and lakes and end up in coastal and marine environments, where they can potentially threaten the health of aquatic ecosystems, and therefore, humans (glassmeyer et al., 2017). finally, intensive agriculture is responsible for the degradation of the quality of surfaceand groundwater, including contamination from pesticides and nitrates (di guardo and finizio, 2016; di guardo et al., 2017). the degradation or even disappearance of aquatic environments put at extreme risk a wide range of important ecosystem services, negatively affecting our well-being and threating our economies. in this context, eu policies promote an integrated framework for water resource management, drawing on european experience of managing river basins and transboundary rivers, and a whole range of european approaches to water and sanitation services. in the past decades, the increasing water contamination levels led the european commission to adopt a series of directives to regulate the over-exploitation of water resources and water pollution (i.e., the eu ‘water framework directive’, or wfd, released on 23 october 2000 http://ec.europa. eu/environment/water/water-framework/index_en. html; the eu ‘marine strategy framework directive’, or ‘msfw’, released on 17 june 2008 https://water.europa.eu/marine/ policy/marine-strategy-framework-directive; the eu ‘maritime spatial planning directive’, or mspd, released on 16 january 2015 https://ec.europa.eu/maritimeaffairs/ publications/maritime-spatial-planning-directive_en/). with these directives, the eu has developed an integrated modelling framework that links environment-use, hydrological and resource-efficiency models to evaluate different scenarios and policy options in terms of efficiency and cost-effectiveness. however, scientists are being more and more frequently questioning the opportunity to update these above-mentioned strategies to keep pace with the continuous conceptual, methodological and technological advancements involving water research, starting from highfrequency monitoring itself (brack et al., 2017; cianelli et al., 2017; ruggiero et al., 2018; buttigieg et al., 2018). in consequence of the present climatic changes driven by carbon dioxide accumulation in the atmosphere, the transition to a low-carbon economy is also one of the key priorities of the european union. essential to achieving this goal is the large-scale deployment of energy technologies with a low carbon-footprint. as for temperate geographical regions, while relatively scanty water quantities, in respect to the overall freshwater budget of these areas, are consumed in renewable energy operations, carbon sequestration and storage practices would require higher amounts of water for cooling and other processes than conventional fossil fuel plants of similar capacities (lal, 2004). measuring the impact of such technologies on water resources is one of the major contribution from the joint research center of the european commission (https://ec.europa.eu/info/ departments/joint-research-centre_en) to the eu’s blueprint to safeguard europe’s water resources and the european energy roadmap to 2050. moreover, extreme climate-change-driven water-related events, such as droughts and floods, can have enormous environmental, social and economic impacts (easterling et al., 2000; neniston et al., 2007). the eu jrc is being developing early warning, monitoring and damage assessment systems for weather-driven natural hazards, since local and transnational disasters have demonstrated that efficient risk management is of paramount importance. jrc also develops tools for assessing the effects of climate and land-use change, risk mapping and adaptation to increasingly frequent extreme events across europe and worldwide (jrc, 2014; thielen del pozo et al., 2015). in light of the deep integration between aquatic science and socio-economic-political issues, it is therefore necessary that the scientific community makes the whole society aware of the results of scientific research on these issues and of the need to intensify the study of the mechanisms that underlie the ongoing changes in aquatic ecosystems (buytaert et al., 2014; albrecht, 2016; hall et al., 2016; jacobs et al., 2016). to help meeting this need, within the 23rd congress of the italian association of oceanology and limnology (aiol; http://www. aiol.info/), entitled “functioning, alteration and recovery of aquatic ecosystems: the aquatic sciences to understand global change and to make the citizens aware of it” (cagliari, italy, 26-29 september 2017), all participants, among which some of the leading national experts in the field of aquatic sciences, were invited to give their contribution in assessing a set of actions needed to achieve an adaptive and proactive management of changes that the aquatic sciences are going to face in the next five years, in order to match societal needs. the 2017 aiol congress was organized to take stock of the most recent research results on structure and functions of terrestrial and marine aquatic ecosystems and on the importance of recovering their functions and biodiversity. the results presented during the congress no nco mm er cia l u se on ly present and future of aquatic sciences according to aiol community 21 aimed providing inputs to implement ecosystem conservation practices, to develop virtuous and sustainable mechanisms to adapt the exploitation of natural resources to environmental and climatic changes, and to plan active recovery actions. during the congress, to bring out the different sensitivities, experiences and competences, a specific workshop was organized where a first general discussion about the future of aquatic sciences in italy was started. then, to involve as many aiol members as possible, at the end of the workshop, a questionnaire consisting of 12 questions organized in four thematic clusters was subsequently circulated online among the whole aiol community. in this paper, the results of the survey are presented and discussed as a contribution to the identification of the main research areas and topics that will require financial investments in the next coming years. methods the questionnaire (tab. 1) was designed as to allow aiol members to participate, regardless of the specific disciplines or fields of interest they dealt with, and to provide their own point of view regarding different aspects of the science-society relationship. furthermore, the survey aimed at understanding the perception that scientists have of themselves, i.e., about their role within the society and how they would like to be envisioned by it. the survey was composed of four clusters of questions (tab. 1). the first one aimed at identifying the main and most current themes of scientific research related to the study of aquatic ecosystems, both marine and freshwater (tab. 1; cluster 1). a maximum of five possible answers was allowed for each question. the all set of given answers to this cluster is available in the supplementary tab. 1. the objective of the second cluster was to evaluate how the scientific community saw the role of its research within the society (tab. 1; cluster 2). eight possible strengths and eight possible weaknesses of scientific knowledge, and their evaluation in terms of relative importance, i.e. fundamental/important/not important/ insignificant, were proposed (supplementary tab. 2). the third cluster aimed at evaluating the perception of the scientific community on social and economic challenges fueled by the current environmental changes stemming from the increasing vulnerability of aquatic environments (tab. 1; cluster 3). a series of possible answers has been proposed, eventually adding an empty field for free answers (supplementary tab. 3). the fourth cluster was used to identify the main research fields that should deserve more attention by the society, according to the aiol community, and the motivation for the given choices. (tab. 1; cluster 4). for each question, a maximum of five possible answers were allowed (supplementary tab. 4). the answers received for clusters 1, 3 and 4 were grouped into sub-groups with a thematic affinity to give a synthetic view of them and to be able to make comparisons among them. since some answers dealt with topics that were across two or more issues, we chose to assign them to the group it looked like to us more coherent with the given answer, and the results of each question were represented by a different pie chart. thirty research scientists attended the congress workshop in cagliari and the completed online questionnaires were thirty, too. results and discussion cluster 1. where are aquatic sciences today? the answers to this first group of questions were very numerous and varied in terms of highlighted issues. some of them were very general (e.g., “climate change”) other more detailed (e.g., “the ecological stoichiometry theory tab. 1. structure of the survey submitted to the aiol scientific community. cluster 1 question 1 what are nowadays the main scientific research topics related to the study of aquatic ecosystems? question 2 which topics, if any, specifically concern marine ecosystems? question 3 which concern only transitional ecosystems? question 4 and which concern freshwater environment? cluster 2 question 5 which are the strengths of scientific knowledge on these issues, and how do you evaluate their relative importance? question 6 which instead are its weaknesses? cluster 3 question 7 which are the social challenges, among those listed below, that will derive from the increasing vulnerability of european and global aquatic environments? question 8 which instead are the economic challenges? cluster 4 question 9 in which fields of basic scientific research should economic and financial resources be invested in the next 5 years? question 10 in which fields of applied scientific research? question 11 in which other specific fields of research linked to land management? question 12 can you list the reasons why, according to you, all these research fields should be funded? for all the questions except for n. 5 and 6, we asked to list or choose a maximum of five topics/items. no nco mm er cia l u se on ly f. alvisi and d. d’alelio22 bridging food webs, ecosystem metabolism and biogeochemistry”) or complex (e.g., “the study of hydromorphological and chemical modifications of lake environments at high altitudes as indicators of climate changes”). therefore, to compare them among each other and with the subsequent questions/clusters, we grouped answers in the following nine subjects (fig. 1a-d): i) life sciences, including biodiversity and ecology issues; ii) climatic sciences and iii) earth sciences, concerning mainly basic research fields and some applied sciences; (iv) toxicology was chosen for answers dealing with environmental health; v) sustainable development, vi) recovery and restoration, vii) protection and conservation included answers specifically addressed to these topics; viii) spatial planning included specific mentions to it as well as the development of useful applications such as modelling, monitoring and shared databases; ix) innovation grouped only answers specifically addressed to it. a huge set of answers mainly concerned the characterization and quantification of the physical, chemical and ecological effects of global changes, in particular the effects of climate change (e.g., global warming, desertification) on aquatic ecosystems and biological communities, as well as on biodiversity and functioning of aquatic ecosystems, and the extent of impacts in terms of ecosystem responses to such changes, the impact on the water cycle and on water resources (e.g., loss of resources, impoverishment and degradation of water quality, changes in the water cycle), and the study of water bodies as environmental indicators of these changes (e.g., water level fluctuation, coastal erosion, salt intrusion, etc.). a second set of answers dealt with the study of biodiversity of aquatic environments and its modifications. it was suggested as a crucial present issue with a focus on its progressive impoverishment due to the strong global anthropogenic pressure in land use and resource exploitation, which led to the decrease or even disappearance of habitats and species of great ecological value. the global scale of these phenomena as well as their acceleration arose a strong interest of the scientific community: as the aiol scientists suggest, attention to the study of the impact of global changes on the species of commercial value (including overfished stocks), the availability of fisheries resources and the possible answers in terms of production to meet the demands of global markets (e.g. aquaculture, fish and mussel farming, etc.) must be paid. moreover, the appearance or reappearance of potentially invasive alien species due to poorly managed introductions and fish restocking, and the increasingly frequent appearance or spread of toxic or harmful species (e.g, dangerous algal blooms, jellyfish), as a result of physical, chemical and ecological alterations of terrestrial and marine water bodies, have become themes of strong scientific and socio-economic interest (boero et al., 2016; hamilton et al., 2014; meriluoto et al., 2017a; wells and bengt, 2018). this was confirmed by our survey’s answers. at the same time, the aiol community also suggested the need to know more and better the morphology and the geological characteristics of the basin floor of both marine (transitional, coastal and deep) and terrestrial (lakes and rivers) water bodies, as a crucial knowledge needed to better link abiotic and biotic factors in a holistic approach to the comprehension of aquatic ecosystem functioning. moreover, traditional fields of research more closely related to the assessment of water quality and of lasting changes induced by water use in production processes and in urban and civil environments (e.g., wastewater, industrial discharges, dispersed sources, groundwater pollution) were also pointed out. these issues have been gradually integrated by studies focused on understanding old and new phenomena related with them, such as eutrophication (e.g., use of fertilizers, sewage discharges; schindler, 2006), acidification (e.g., increase in co2 in atmosphere; verspagen et al., 2014), pollution (e.g., microplastics and marine litter; duis and coors, 2016) and contamination (e.g., contaminants of emerging concern = cecs). as suggested by our survey, the assessment of the effects of these severe and widespread alterations (e.g., toxicology studies) has proven to be increasingly necessary and crucial to respond effectively and incisively to the increase in diseases and allergies and other threads on human and environmental health. according to our respondents, the aquatic science community recently began to focus also on more transversal aspects that put the system, its functioning and relations with other neighboring systems at the center of interests. these latter issues include i) scaling of connectivity in marine and freshwater populations, ii) conservation ecology and spatial planning, iii) ecology, recovery/restoration of degraded habitats (restoration ecology), iv) integration between ecology and economic and social sciences, v) blue economy, and vi) quantification of ecosystem services. the shift towards a so-called “systems-ecology” has led to the design and development of new and increasingly sophisticated tools and technologies for the detection of early impacts, the implementation of interoperable observing systems, the increasingly widespread use of modeling, and the construction of interactive databases. survey results highlighted other increasingly specific and often innovative fields that recently begun to find space and interest: the study of the relationship between biodiversity and ecosystem functioning; no nco mm er cia l u se on ly present and future of aquatic sciences according to aiol community 23 the microbial diversity in the deep sea and deep lakes; the study of viral diversity in the aquatic environment; biological interactions and microbial symbioses; ‘classic’ biogeochemistry, and that modeled with emerging trait-based genomics; ecological stoichiometry theory bridging food webs, ecosystem metabolism and biogeochemistry; the development of strategies to guarantee environmental sustainability (bioprospecting, circular economy, bioremediation); the blue biotechnology, or, the discovery and use of molecules produced by aquatic organisms that contain commercially-exploitable chemically-reactive biomolecules. after comparing general research fields (fig. 1a) with those specifically addressed to marine (fig. 1b), transitional (fig, 1c) or freshwater (fig. 1d) aquatic ecosystems suggested by our community, we notice that (fig. 2): 1) as for climate and life sciences, with emphasis on the biodiversity theme and related issues (see above), the answers are mostly homogeneous and concern all the territorial contexts suggesting the great attention of research scientists to themes pertaining climate change and its impacts on all ecosystems; 2) as for earth sciences, a lower percentage of the general topics emerged if compared to the importance of those specifically regarding transitional environments, which underlines the need for a better understanding of the sensitivity of these latter environments to climate and environmental changes; 3) great attention is also paid to toxicology and environmental health, in particular for marine and lacustrine aquatic environments, rather than for transitional ones, which suggests the strong link between water quality and the quality of derived resources (e.g. fishery, aquaculture, drinking water, tourism, etc.); 4) a general minor interest by aquatic scientists can be seen for the other categories compared to the previous ones and even for the innovation issue, which suggests that the aquatic science community is still strongly oriented to basic scientific themes, though, at the same time, needing a greater integration and interaction with other societal subjects to improve its applicative and managerial attitude. in this new cultural context, even the figure of the aquatic scientist is changing and is assuming multidisciplinary facets and approaches that span from sociology to biology, from physics to chemistry, from engineering to economy. the strong and driven fig. 1. main research fields in the present scientific aquatic sciences community: a) general, b) marine, c) transitional, d) freshwater, according to aiol community. values are in % of received answers. no nco mm er cia l u se on ly f. alvisi and d. d’alelio24 specialization typical of researchers involved in the various disciplines of aquatic sciences, which remains necessary to understand the single steps of complex ecological processes and environmental phenomena, could clash with the need to have a holistic vision, a capacity for synthesis that would allow addressing the various scientific challenges with an eye always focused on socio-economic disciplines and on the global context. this is fostering new needs and approaches converging towards the design of multiand interdisciplinary projects for water monitoring and management (dickey, 2009; tyler et al., 2009; buttigieg et al. 2018). cluster 2. strengths and weaknesses of scientific knowledge the second survey cluster was aimed at assessing the researchers’ perception about the importance of what they do. so, the questions were concerning what the actual strength points of the scientific knowledge on current aquatic science issues are and how do scientists evaluate their relative importance. the obtained answers (fig. 3) give us the idea that most of the proposed items (fig. 3 af) were judged important/fundamental up to 90-93% of the community, except for the last two (fig. 3 g-h) that recorded a slightly lower (73-77%) consensus. the scientific knowledge was considered as the main tool (fig. 3a: fundamental = 70%) to acquire and improve basic knowledge on aquatic systems. an important point was also about the applied knowledge (fig. 3b: important = 60%) seen as a tool to deal with either environmental problems or challenges. nevertheless, the results of the survey suggest that both basic and applied research should be coupled to communication skills (fig. 3d): half of our responding group considered communication as fundamental (47%) to allow sharing aquatic knowledge and expertise with stakeholders aquatic knowledge and expertise, and then to translate these latter in usable, recognizable and exploitable information for the society. equally important (60%), even if judged as not fundamental (34%), are the ability to network inside and outside the scientific community (fig. 3c), the management approach (fig. 3e: 63% and 27%, respectively) as well as the forecasting capacity (fig. 3f: 64% and 30%, respectively). when moving on to socio-economic issues, the perception of the importance of scientific knowledge regarding the general well-being of the society (fig. 3g: 60% and 13%, respectively) and its economic development (fig. 3h: 54% and 23%, respectively) turns out to be even weaker. this suggests that it is not always possible tackling these aspects with the society or that the interconnection between science and society is not straightforward. we can interpret all results pertaining the above-mentioned questions as a sign that scientists’ selfperception is more as being a ‘carrier and transmitter of knowledge’ rather than as an ‘authoritative interlocutor for the society regarding prediction, management and resolution of environmental problems or challenges’. fig. 2. comparison among answers received for different environmental contexts. values are in % of received answers. no nco mm er cia l u se on ly present and future of aquatic sciences according to aiol community 25 when, in turn, we asked about the perception about scientific-knowledge weaknesses (fig. 4), the proposed items were all judged fundamental/important, with up to 97% of consensus given to the lack of adequate policies about public research (fig. 4f). for about one third of the interviews, gaps in knowledge (fig. 4a) and paucity of skills (fig. 4b) were fundamental (33% and 30%, respectively) and for more than a half (53% and 54%, respectively) these were important points of weakness, but there were also 1317% of the respondents who considered these items not important. we can interpret these results as the common and fixed feeling by which scientists assume that new disciplines and new attitudes are always and anyway required to face new scientific challenges and topics. fig. 3. perceived strengths of the scientific knowledge on aquatic science issues and rating of their relative importance. values are in % of received answers. no nco mm er cia l u se on ly f. alvisi and d. d’alelio26 the results of the survey also indicate that for the development of scientific knowledge the lack of scientific tools (fig. 4c) or infrastructures (fig. 4d) are perceived as important (70% and 67% respectively), though not so fundamental as the lack of funding (fig. 4e) and adequate policies about research (fig. 4f). finally, there is a clear perception of a weak connection between science and society, partially due to the low awareness of citizens about the ‘world’ of research (fig. 4g), as well as to the self-referentiality of researchers (fig. 4h). these answers point out the presence of two apparently opposing views on the role of the scientists, which conceal a perennial conflict. on the one hand, some researchers think they are an integral part of society, fig. 4. perceived weaknesses of the scientific knowledge on aquatic science issues and rating of their relative importance. values are in % of received answers. no nco mm er cia l u se on ly present and future of aquatic sciences according to aiol community 27 contributing with their work to increase human communities’ well-being and wealth, and recognize the importance of communication skills and the value of knowledge, but also of networking and managerial skills. at the same time, they are critically aware about the tendency of scientists to be self-referent. conversely, some other researchers think that scientific work must be only focused on acquiring specific knowledge and skills, and that citizen and society should be more interested in research, finance it more and better, to recognize its value and usefulness. paradoxically, this group of scientists does not seem to take into full consideration their citizenship. this condition might derive from the prolonged absence of a clear research policy at national level that would cause a weakening of affinity between scientists and modern societies. cluster 3. perception of socio-economic challenges in a changing scenario to assess whether the scientists can catch the closely approaching social and economic challenges that society is going to meet in the next few years, as well as those predicted in the medium to long-term, we posed two questions concerning challenges potentially linked to the increasing vulnerability of aquatic environments at european and global levels (cluster 3; tab. 1). as for social challenges, the obtained preferences (figs. 5 and 6) pertained some purely objective aspects, with a focus on the shortage of water resources (18%) and drought (5%), and on their direct impacts on land, such as a general environmental degradation (e.g., land and ecosystem degradation =15%; hydrogeological instability =15%, desertification =7%). nevertheless, some of the main social aspects strictly linked to water shortage, such as ensuring health and well-being (e.g., urban settlements development, coastal and internal areas habitability, tourist-flows management, migratory-flows increase), were pointed out as equally important (29%, fig. 6). lower preferences (fig. 6) were granted to the challenges connected with the protection of natural and cultural heritage (7%) and to the need of more research and development (r&d) (4%). even if these results were partially affected by the number and nature of proposed answers (see groups of fig. 5), they testify scientists’ clear perception that land degradation is the main social challenge we must face, since its effects are increasingly affecting the citizens’ well-being and health. in particular, the ability to manage large flows of people caused by global environmental changes, the so-called ‘environmental migrants’, is already now, and will increasingly constitute a crucial challenge to ensure political stability, social equity and economic sustainability at national and international level (science for environment policy, 2015). fig. 5. social challenges that will result from the increasing vulnerability of european and global aquatic environments and rating of their relative importance. colors refer to groups shown in fig. 6. values are in % of received answers. no nco mm er cia l u se on ly f. alvisi and d. d’alelio28 concerning the perception of the link between the increasing vulnerability of aquatic environments and derived economic challenges, the answers highlighted the need to adopt a diffuse sustainable approach to all the main economic issues (figs. 7 and 8), such as food production (16%), clean energy (10%), raw material (10%), fishing (6%) and agriculture (5%), as well as a general intervention to cope with climatic changes (13%) (fig. 7). health and wellness of citizens and society as a whole (24%; fig. 8) have been again identified as crucial targets for the next years, including the direct management of health and wellness (10%), demographic changes (7%), food security (5%), and supportive, innovative and safe communities (2%) aspects (fig. 7). as seen for social challenges, also in this case specific knowledge or skills challenges appear to be embedded one inside the other more than seen as separate issues (fig. 8: r&d = 1.4%). this last result suggests that the aiol community seems to be self-confident that what has already being done or could be done soon in terms of r&d (0.7%), and coordination and collaboration between national and international bodies (0.7%), does not represent a challenge, but rather an established practice which only needs greater acknowledgement by means of economic and political support, as previously stated above. cluster 4. which aquatic science topics to finance over the next five years and why? to understand what, among the many disciplines of aquatic sciences, the research priority fields to invest in should be in the next few years, we submitted two questions aimed at distinguishing among basic and applied research funding priorities, and a further one specifically addressed to land management issues (tab. 1). fig. 6. social challenges grouped by main topic and relative preferences. values are in % of received answers. fig. 7. economic challenges that will result from the increasing vulnerability of european and global aquatic environments and relative preferences. colors refer to groups shown in figure 8. values are in % of received answers. no nco mm er cia l u se on ly present and future of aquatic sciences according to aiol community 29 among basic research topics, life sciences have obtained 43% of the preferences, followed by earth (12%) and climate sciences (9%), with a specific emphasis about protection and conservation (3%), but without any mention about recovery and restoration (0%) (fig. 9). in addition to them, a request for a special attention to environmental health, and more properly to water toxicology, was expressed by the 12% of the sample, as well as the need to see a widespread development of a better spatial planning (9%) with a special attention to economic/ecological/ecosystem efficiency and sustainability (8%) (fig. 9). to give an idea of the complexity and variety of basic research topics that were proposed for short-term funding within the survey, we report the following items, among others, declared by the respondent. as concerning life sciences, the main mentioned topics were: marine and freshwater habitat characterization, biodiversity and ecology of marine and freshwater organisms, bio-ecology of commercial species, biodiversity assessment and conservation, alien species, (microbial) ecology, ecological and ecosystem processes, ecosystem functioning, relationship between biodiversity and ecosystem functioning, mechanisms of ecosystem changes over time, biomimicry, the relationship between ecology and evolution, taxonomy, genetics, metagenomics, metabolomics, altogether accounting for the 43%. in the earth sciences sector, topics to be financed should be, among others: the response to global changes in terms of hydrological alterations, connectivity and continuity of freshwater systems, water quantity and quality, biogeochemical cycles, and a general better knowledge of lacustrine/marine geological history and evolution. as for climate sciences were concerned, specific funds should be dedicated to deepen our understanding fig. 8. economic challenges grouped by main topic and relative preferences. values are in % of received answers. fig. 9. main basic research issues where to invest in in the next five years, according to aiol community. values are in % of received answers. no nco mm er cia l u se on ly f. alvisi and d. d’alelio30 about climate changes and on their impact on ecosystems, and to develop new approaches and strategies in order to counteract such changes (e.g., blue carbon sequestration). as concerning the environmental management, some specific topics to be financed should be devoted to a widespread monitoring of environmental quality, with the collection of high-frequency data, building of permanent infrastructures for the observation of aquatic systems, shared and interoperable databases. the aim is to develop a better spatial planning, by including protection and management of ecosystems, but also sustainable use of biotic and abiotic resources, by improving the efficiency of water use, by promoting a closer connection with sociology and new economy, and by investing in natural capital. a great concern was also expressed about the need to concentrate research funding on a general improvement of water quality, such as focusing on the relationships between health and pollution, on the reduction of contamination and marine litter, on the effects of pollution on the trophic webs, on the effects of cecs on structure and composition of aquatic biocoenosis, on the role of antibiotic substances, on the development of rapid microtoxin analysis to evaluate risks and effects of toxic algal blooms, etc. concerning applied-research topics to be funded over the next 5 years, the results of the survey (fig. 10) clearly point out to a wider concern on the need to financially support a transition towards a sustainable development approach (26%). we would like to point out that in the group “sustainable development” (fig. 10) we included only the survey answers that contained specific indications about “sustainability”. however, most of the given answers, grouped in the other sub-clusters of figure 10, contained references and links to it, suggesting that this is felt as a major issue. applied research, according to the aiol community, should be carried forward by means of a widespread use of spatial planning (14%), based on monitoring and modelling of ecosystems characteristics, processes and functioning, run by permanent infrastructures for the observation of aquatic systems, in order to build shared and interoperable databases through the identification of reliable indicators and a common language to allow a more effective and efficient integration of ecology within economic and social sciences. dedicated studies on recovery and restoration (13%), as well as protection and conservation (5%) of aquatic ecosystems, are fundamental pillars of this approach (fig. 10), which should take advantage of innovative methods and techniques (14%). an increasingly deeper and widespread connection between research and productive sectors that should be fig. 10. priority fields to invest in the next five years across the applied disciplines of aquatic sciences according to aiol community. values are in % of received answers. no nco mm er cia l u se on ly present and future of aquatic sciences according to aiol community 31 funded within the framework of r&d collaboration projects is therefore highly desirable. in addition, the recovery and restoration group (fig. 10) included mainly: species responses in terms of production of secondary metabolites toxic and/or useful for humans, reduction of chemical and physical pollution/contamination in aquatic ecosystems, surface and groundwater depollution/depuration, containment of (micro)plastics and of emerging pollutant (cecs) diffusion, increase of treatment plants, ecotoxicology vs. oncology studies. as for protection and conservation (fig. 10), among the survey answers, we found issues such as the establishment of a network of protected areas, actions for biodiversity conservation, applied ecology and evolution, modeling of biogeochemical cycles, definition of ecological standard for water quality and quantity, effect of biodiversity loss on ecosystem functioning, microbiology, astrobiology. other examples of “hot” topics that should deserve the attention of national and international funding agencies, according to survey answers were: adaptation vs. mitigation measures to climate change, use of aquatic environments for the development of clean energy, process of reuse of natural resources and their protection, improved use of water in food and goods production, sustainable agriculture, aquaculture and fisheries, measures and technologies to preserve human and animal wellbeing, development of new technologies and platforms with low environmental impact, study of the impact on the environmental and architectural quality of cities by the widespread application of green techniques. the number and complexity of the given answers confirms the multidisciplinary and interdisciplinary nature of aquatic science and the difficulty to focus on few key issues. the subsequent question we asked, about research topics addressed only to environmental management (tab. 1), was therefore aimed at encouraging the aiol community to deeply focus on this multidisciplinary and interdisciplinary nature of aquatic science to find connections and interactions among basic and applied research. the answers they provided (fig. 11), grouped as the previous ones to allow a more precise comparison, suggest that the aiol community has clear in mind the need to move toward a widespread sustainable approach (sustainable development 18%), that must be based on a better and severe scientific approach to land management (earth sciences 18%) and improvement of environmental quality (toxicology = 8%). the aim is to adopt a wise spatial planning strategy (18%), as a diffused and shared management tool, by also taking into consideration recovery and restoration (10%) as well as preservation and fig. 11. priority fields to invest in the next five years across the disciplines of aquatic sciences applied to environmental management, according to aiol community (% value of received answers). no nco mm er cia l u se on ly f. alvisi and d. d’alelio32 conservation (6%) actions. all this can and must be a strong push towards innovation (16%), not only in terms of instrumentations and technology, but also and above all in the way of thinking about the management of the territory and natural resources. given all that, why will civil society finance all this? what are the possible risks if we do not adopt this strategy or delay these interventions in favor of a more widespread and efficient research applied to land management? we provided a series of possible answers and an empty field for free answers. survey answers (fig. 12) suggested an awareness about the urgency to prevent a further deterioration of aquatic and terrestrial ecosystems and wetlands (15%). but they were also focused on the need for a mediumto long-term vision (12%), in which both critical approach and scientific knowledge are at the basis of the elaboration of sustainable management plans (11%) to improve the status of aquatic ecosystems (10%). as a final remark, from the list of given answers (fig. 12) we notice that they are often coupled twofold: one of them reflects a purely scientific motivation while the other contains a social implication, as if to testify the ‘professional vs. personal’ dual nature of each research scientist, who has to express her/himself on issues in which emotional involvement can play a fundamental role too. some examples are (fig. 12): i) to involve local communities in land participatory management (9%) vs. to advance scientific knowledge on these topics (9%); ii) to protect the status of aquatic and terrestrial ecosystems and wetlands (6%) vs. to contribute to the construction of a shared system of rules based on the results of scientific research (6%); iii) to promote and contribute to training new professionals (5%) vs. to contribute to a widespread cultural growth of society (5%). conclusions the results of this survey allowed us to set the basis for a potential list of the priorities that funding agencies should include in their economic and financial planning of the next five years: 1) to promote and support sustainable food production (e.g., agriculture, aquaculture and fish farming), develop integrated and sustainable systems in aquaculture, boost the control of the food supplychain, etc.; 2) to promote and diffuse the spatial planning and integrated ecosystem-based management approach of maritime and terrestrial space by adopting and fig. 12. some reasons why the society should invest economic and financial resources in the next five years in these research fields, according to survey results. values are in % of received answers. no nco mm er cia l u se on ly present and future of aquatic sciences according to aiol community 33 supporting positive actions for i) participatory democracy, ii) environmental control of urban settlements and landscape, iii) green and blue economy including energy, food supply and use of water resources, etc.; 3) to develop recovery/remediation plans for contaminated sites (e.g., by means of bioremediation), as well as intervention plans to i) counteract environmental degradation (terrestrial and marine) and hydrogeological instability, ii) assess cecs impact on ecosystems, iii) promote biological control of organisms harmful to agriculture and forest ecosystems, iv) limit the spread of ballast waters, etc.; 4) to promote conservation of ecosystems by i) assessing their conservation status, first of all the water quality, ii) evaluating and quantifying their ecosystem services, iii) applying a watershed ecology, iv) identifying of areas to be protected, even in deep marine environments, etc.; 5) to foster the technological development of sustainable and integrated tools and procedures for environmental monitoring aimed at i) ensuring ecosystem functionality, ii) identifying and minimizing the impact of anthropogenic pressures on aquatic systems, iii) developing recovering techniques for compromised aquatic environments in order to recover their ecosystem services, iv) applying realtime monitoring techniques (e.g., remote sensing) to water bodies in urban and peri-urban areas (e.g., mega-cities), etc.; 6) to develop a better forecasting capacity, particularly of extreme events, by i) implementing long-term research networks, ii) improving reliable predictive techniques of pressures entities, iii) developing modeling-methods based on geo-referenced systems in order to plan management and restoration actions of ecological quality, iv) implementing early warning systems, etc.; 7) to support training courses for professionals and a learning processes extended to a wider society, in order to i) promote new jobs, ii) further encourage interdisciplinary and multidisciplinary research, iii) achieve a more effective transfer of knowledge, iv) involve citizens in research (i.e., citizen science), v) promote social inclusion and culture, vi) evaluate ecosystem health, and finally vii) gain mutual understanding and interoperability between science and society. acknowledgments we wish to thank all the members of the associazione italiana di oceanologia e limnologia (aiol) who participated to the cagliari workshop and to the survey, and antonio pusceddu and nico salmaso for their support and the confidence placed in us. references albrecht j, 2016. legal framework and criteria for effectively coordinating public participation under the floods directive and water framework directive: european requirements and german transposition. environ. sci. pol. 55:368-375. alpine convention, 2009. water and water management issues report on the state of the alps. alpine signals, special edition 2. innsbruck. beniston m, stephenson db, christensen ob, ferro ca, frei c, goyette s, halsnaes k, holt t, jylhä k, koffi b, palutikof j, 2007. future extreme events in european climate: an exploration of regional climate model projections. clim. change 81:71-95. bertoni r, 2006. [laghi e scienza introduzione alla limnologia].[book in italian]. aracne editrice, roma. boero f, bonsdorff e, 2007. a conceptual framework for marine biodiversity and ecosysm functioning. mar. ecol. 28: 34-145. boero f, brotz l, gibbons mj, piraino s, zampardi s, 2016. impacts and effects of ocean warming on jellyfish. gland, switzerland, iucn: 213-237. bogdan a, istudor n, gruia r, tobă gf, bulz n, gâf-deac i, chelmu s, găvan c, prică i and paşalău c, 2014. new holistic approach of bioeconomics and ecoeconomics theories, practical bridging from the green economy to blue economy, trough new integrated and innovative paradigm about “bio-eco-geo-economy”. proc. econ. fin. 8:83-90. brack w, dulio v, ågerstrand m, allan i, altenburger r, brinkmann m, bunke d, burgess rm, cousins i, escher bi, hernández fj, hewitt lm, hilscherová k, hollender j, hollert h, kase r, klauer b, lindim c, herráez dl, miège c, munthe j, o’toole s, posthuma l, rüdel h, schäfer rb, sengl m, smedes f, van de meent d, van den brink pj, van gils j, van wezel ap, vethaak ad, vermeirssen e, von der ohe pc, vrana b, 2017. towards the review of the european union water framework management of chemical contamination in european surface water resources. sci. total environ. 576:720-737. brock td, 1985. long-term change in lake mendota. in: a eutrophic lake. ecological studies (analysis and synthesis), 55. springer, new york. doi.org/10.1007/978-1-44198700-6_9 buttigieg pl, fadeev e, bienhold c, hehemann l, offre p, and boetius a, 2018. marine microbes in 4d using time series observation to assess the dynamics of the ocean microbiome and its links to ocean health. curr. opin. microbiol. 43:169-185. buytaert w, zulkafli z, grainger s, acosta l, alemie tc, bastiaensen j, de bièvre b, bhusal j, clark j, dewulf a, foggin m, hannah dm, hergarten c, isaeva a, karpouzoglou t, pandeya b, paudel d, sharma k, steenhuis t, tilahun s, van hecken g, zhumanova m, 2014. citizen science in hydrology and water resources: opportunities for knowledge generation, ecosystem service management, and sustainable development. front. earth. sci. 2:26. no nco mm er cia l u se on ly f. alvisi and d. d’alelio34 cianelli d, d’alelio d, uttieri m, sarno d, zingone a, zambianchi e, ribera d’alcalà m, 2017. disentangling physical and biological drivers of phytoplankton dynamics in a coastal system. sci. rep. 7:15868. crain cm, kroeker k, halpern bs, 2011. interactive and cumulative effects of multiple human stressors in marine systems. ecol. lett. 11:1304-1315. d’alelio d, libralato s, wyatt t, ribera d’alcalà m, 2016. ecological-network models link diversity, structure and function in the plankton food-web. sci. rep. 6:21806. di guardo a, finizio a, 2016. a moni-modelling approach to manage groundwater risk to pesticide leaching at regional scale. sci. total environ. 545/546:200-209. di guardo a, terzaghi e, raspa g, borin s, mapelli f, chouaia b, zanardini e, morosini c, colombo a, fattore e, davoli e, armiraglio s, sale vm, anelli s, nastasio p, 2017. differentiating current and past pcb and pcdd/f sources: the role of a large contaminated soil site in an industrialized city area. environ. poll. 223:367-375. dornburg v, faaij apc, verweij pa, banse m. diepen kv, keulen hv, langeveld h, meeusen m, ven gvd, wester f, alkemade r, 2008. biomass assessment: assessment of global biomass potentials and their links to food, water, biodiversity, energy demand and economy: inventory and analysis of existing studies: supporting document (no. 500102 014), mnp. dickey t, 2009. progress in multi-disciplinary sensing of the 4dimensional ocean. in: w. hou (ed.), ocean sensing and monitoring. spie proceedings 7317:731702. duis k, coors a, 2016. microplastics in the aquatic and terrestrial environment: sources (with a specific focus on personal care products), fate and effects. environ. sci. eur. 28:2. easterling dr, meehl ga, parmesan c, changnon sa, karl tr, mearns lo, 2000. climate extremes: observations, modeling, and impacts. science 289:2068-2074. fattore e, davoli e, armiraglio s, sale v m, anelli s, nastasio p, 2017. differentiating current and past pcb and pcdd/f sources: the role of a large contaminated soil site in an industrialized city area. environ. poll. 223:367-375. gamfeldt l, lefcheck js, byrnes jek, cardinale bj, duffy je, griffin jn, 2015. marine biodiversity and ecosystem functioning: what’s known and what’s next? oikos 124:252265. glassmeyer st, furlong et, kolpin dw, batt al, benson r, boone js, conerly o, donohue mj, king dn, kostich ms, mash he, pfaller sl, schenck km, simmons je, varughese ea, vesper sj, villegas en, wilson vs, 2017. nationwide reconnaissance of contaminants of emerging concern in source and treated drinking waters of the united states. sci. total environ. 581/582:909-922. hall dm, gilbertz sj, anderson mb, ward lc, 2016. beyond “buy-in”: designing citizen participation in water planning as research. j. clean prod. 133: 25-734. hamilton dp, wood sa, dietrich dr, puddick j, 2014. costs of harmful blooms of freshwater cyanobacteria, p. 245-256. in: n.k. sharma, a.k. rai and l.j. stal (eds.), cyanobacteria: an economic perspective. j. wiley & sons, chichester. howarth r, chan f, conley dj, garnier j, doney sc, marino r, billen g, 2011. coupled biogeochemical cycles: eutrophication and hypoxia in temperate estuaries and coastal marine ecosystems. front. ecol. environ. 9:18-26. jacobs k, lebel l, buizer j, addams l, matson p, mccullough e, garden p, saliba g, finan t, 2016. linking knowledge with action in the pursuit of sustainable water-resources management. p. natl. acad. sci. usa 113:4591-4596. mortara b, gonzalez verdesoto e, barry g, 2014. science for disaster risk reduction. jrc thematic report, publications office of the european union. doi: 10.2788/65084. kallf j, 2002. limnology : inland water ecosystems. prenticehall, inc., upper saddle river: 592 pp. lal r, 2004. soil carbon sequestration impacts on global climate change and food security. science 304:1623-1627. likens ge, 2009. encyclopedia of inland waters. 1. academic press: 2250 pp. manca m, calderoni a, mosello r, 1992. limnological research in lago maggiore: studies on hydrochemistry and plankton. mem. ist. it. idrobiol. 50: 71-200. meriluoto j, blaha l, bojadzija g, bormans m, brient l, codd ga, drobac d, faassen ej, fastner j, hiskia a, ibelings bw, kaloudis t, kokocinski m, kurmayer r, pantelić d, quesada a, salmaso n, tokodi n, triantis tm, visser pm, svirčev z, 2017a. toxic cyanobacteria and cyanotoxins in european waters recent progress achieved through the cyanocost action and challenges for further research. adv. ocean. limnol. 8:161-178. doi: 10.4081/aiol.2017.6429 meriluoto j, spoof l, codd ga, 2017b. handbook on cyanobacterial monitoring and cyanotoxin analysis. wiley: 576 pp. morabito g, mazzocchi mg, salmaso n, zingone a, bergami c, flaim g, accoroni s, basset a, bastianini m, belmonte g, bernardi aubry f, bertani i, bresciani m, buzzi f, cabrini m, camatti e, caroppo c, cataletto b, castellano m, del negro p, de olazabal a, di capua i, elia ac, fornasaro d, giallain m, grilli f, leoni b, lipizer m, longobardi l, ludovisi a, lugliè a, manca m, margiotta f, mariani ma, marini m, marzocchi m, obertegger u, oggioni a, padedda bm, pansera m, piscia r, povero p, pulina s, romagnoli t, rosati i, rossetti g, rubino f, sarno d, satta ct, sechi n, stanca e, tirelli v, totti c, pugnetti a, 2018. plankton dynamics across the freshwater, transitional and marine research sites of the lter-italy network. patterns, fluctuations, drivers. sci. tot. environ. 627: 73-387. mosello r, calderoni a, de bernardi r, 1997. le indagini sulla evoluzione dei laghi profondi sudalpini svolte dal c.n.r. istituto italiano di idrobiologia. docum. ist. it. idrobiol. 61:19-32. pauli ga, 2010. the blue economy: 10 years, 100 innovations, 100 million jobs. paradigm publications: 308 pp. pearce f, 2006. when the rivers run dry: water – the defining crisis of the twenty-first century. beacon press: 336 pp. peters rh, 1990. pathologies in limnology. mem. ist. it. idriobiol. 47:181-217. postel s, carpenter s, 1997. freshwater ecosystem services., p. 195-214. in: g. daily (ed.), nature’s services: societal dependence on natural ecosystems. island press, washington, dc. reichman o j, jones mb, schildhauer mp, 2011. challenges no nco mm er cia l u se on ly present and future of aquatic sciences according to aiol community 35 and opportunities of open data in ecology. science 331:703705. rösch r, baer j, brinker a, 2017. impact of the invasive threespined stickleback (gasterosteus aculeatus) on relative abundance and growth of native pelagic whitefish (coregonus wartmanni) in upper lake constance. hydrobiologia 1-12. doi: 10.1007/s10750-017-3479-6 ruggiero mv, d’alelio d, ferrante mi, santoro m, vitale l, procaccini g, montresor m, 2018. clonal expansion behind a marine diatom bloom. isme j. 12:463-472. schindler dw, 2006. recent advances in the understanding and management of eutrophication. limnol. oceanogr. 51: 356-363. science for environment policy, 2015. migration in response to environmental change. science communication unit (ed.), uwe, bristol 51: 16 pp. silver jj, gray nj, campbell lm, fairbanks lw, gruby rl, 2015. blue economy and competing discourses in international oceans governance. j. environ. dev. 24(2):135160. thielen del pozo j, thiemig v, pappenberger f, revilla-romero b, salamon p, de groeve t, hirpa f, 2015. the benefit of continental flood early warning systems to reduce the impact of flood disasters. eur 27533 en; doi:10.2788/46941 tyler an, hunter pd, carvalho l, codd ga, elliott ja, ferguson ca, hanley nd, hopkins dw, maberly s c, mearns kj, scott em, 2009. strategies for monitoring and managing mass populations of toxic cyanobacteria in recreational waters: a multi-interdisciplinary approach. environ. health 8:s11. verspagen jmhh, van de waal db, finke jf, visser pm, van donk e, huisman j, 2014. rising co2 levels will intensify phytoplankton blooms in eutrophic and hypertrophic lakes. plos one 9:e104325. vilà m, basnou c, pyšek p, josefsson m, genovesi, p, gollasch s, nentwig w, olenin s, roques a, roy d, hulme pe, 2010. how well do we understand the impacts of alien species on ecosystem services? a pan-european, cross-taxa assessment. front. ecol. environ. 8:135-144. water framework directive, 2000. directive 2000/60/ec of the european parliament and of the council of 23 october 2000 establishing a framework for community action in the field of water policy. official journal of the european parliament. wells ml, karlson b, 2018. harmful algal blooms in a changing ocean. in: p.m. glibert, e. berdalet, m.a. burford, g.c. pitcher, m. zhou (eds.), global ecology and oceanography of harmful algal blooms. springer: 77-90. wetzel rg, 2001. limnology. lake and river ecosystems. academic press, san diego, usa: 1006 pp. williams wd, 2001. anthropogenic salinisation of inland waters, p. 329-337. in: j.m. melack, r. jellison, d.b. herbst (eds.), saline lakes. developments in hydrobiology. springer, dordrecht. no nco mm er cia l u se on ly layout 1 introduction bottom trawling, along with dredging and dumping, represents one among the most severe physical disturbances generated by human activities at sea (thrush and dayton, 2002). typically, bottom trawling is carried out using heavy ground ropes and chains to drive fish and crustaceans from the seabed into nets (johnson et al., 2015). trawling is carried out on many types of grounds, from shallow waters down to the deep continental margins (puig et al., 2012), by small and large vessels, and for a wide range of target species, including fish and crustaceans (hinz et al., 2009). these characteristics make bottom trawling one of the preferred methods of industrial fisheries worldwide but, at the same time, one among the human activities at sea most impacting highly vulnerable benthic marine ecosystems (e.g., cold-water corals or coralligenous bottoms; fowler, 2003; althaus et al., 2009; bruckner, 2009; heifetz et al., 2009; bongiorni et al., 2010). recently, not irrelevant damages determined by bottom trawling have been also documented on benthic communities and processes in deep-sea soft bottoms (pusceddu et al., 2014). previous investigations carried out in shallow marine ecosystems have revealed that bottom trawling can generate a plethora of direct and indirect effects on pelagic and benthic environments and biota (kaiser, 1998; smith et al., 2003; thrush and dayton, 2002; queiros et al., 2006; hiddink et al., 2007; smith et al., 2013). as well as having a direct impact on the stocks of the target species and the by-catch, bottom trawling can also alter the structure and physico-chemical characteristics of the trawled sediment and of the overlying water column (jennings et al., 2001; smith et al., 2003; pusceddu et al., 2005b, 2005c; puig et al., 2012). in the pelagic realm, bottom trawling can increase turbidity, internal nutrient loads, oxygen consumption, and possibly enhance phytoplankton primary production (riemann and hoffmann, 1991; palanques et al., 2001; durrieu de madron et al., 2005). advances in oceanography and limnology, 2015; 6(1/2): 21-32 original article doi: 10.4081/aiol.2015.5448 quantity and biochemical composition of particulate organic matter in a highly trawled area (thermaikos gulf, eastern mediterranean sea) antonio pusceddu,1* silvia bianchelli,2 roberto danovaro2 1dipartimento di scienze della vita e dell’ambiente, università degli studi di cagliari, via fiorelli 1, 09126 cagliari; 2dipartimento di scienze della vita e dell’ambiente, università politecnica delle marche, via brecce bianche, 60131 ancona, italy *corresponding author: apusceddu@unica.it abstract bottom trawling represents nowadays one of the most severe anthropogenic disturbances at sea, and determines large impacts on benthic communities and processes. bottom trawling determines also local sediment resuspension and the effects of the injection of large amounts of surface sediments into the water column have been repeatedly investigated. few studies have assessed the consequences of sediment resuspension caused by bottom trawling on the quantity, biochemical composition and bioavailability of suspended organic particles and how these eventually rival those exerted by natural storms. to provide insights on this poorly addressed issue, we investigated concentrations and biochemical composition of total and enzymatically digestible pools of particulate organic matter (pom) in the thermaikos gulf (mediterranean sea) under calm sea conditions, during intensive trawling activities, and after a severe storm. we show here that sediment resuspension caused by trawling can cause large effects on pom quantity, biochemical composition and bioavailability. both during trawling and after the storm, the relative importance of the carbohydrate pools increased (in the upper water column) and the total lipid concentrations decreased (in the intermediate and bottom layers) when compared to values measured during calm conditions. these results would suggest that bottom trawling could inject in the upper water column pom pools more refractory in nature (e.g., carbohydrates) than those present in calm or after-storm conditions. by contrast, we show also that the bioavailable fraction of biopolymeric c increased significantly during trawling in the upper water column of the shallowest stations and in the bottom water column layer of the deepest ones. these results provide evidence that bottom trawling can influence the overall trophic status of coastal waters, exerting effects similar or stronger than those caused by natural storms, though of variable amplitude depending on the water depth. since bottom trawling is carried out worldwide and natural storms at sea can be frequent and intense, we claim for the need of assessing new adapting management strategies of bottom trawling in order to mitigate the synergistic impacts of anthropogenic and natural sediment resuspension on coastal biogeochemical cycles. key words: particulate organic matter; bottom trawling; eutrophication; mediterranean sea. received: july 2015. accepted: october 2015. no nco mm er cia l u se on ly 22bottom trawling impacts on particulate organic matter bottom trawling can play also a key role in sustaining high productivity on some continental margins by accelerating sedimentary c degradation (polymenakou et al., 2005), nutrient turnover and, thus, enhancing phytoplankton blooms (fanning et al., 1982; christensen, 1989). one of the evident effects of bottom trawling consists in sediment resuspension which generates visible and highly turbid plumes of suspended particles, with concentrations up to several hundred mg l–1 near the seabed (schoellhamer, 1996; durrieu de madron et al., 2005). the injection of large amounts of surface sediments into the water column, particularly when trawling is carried out over soft bottoms (black and parry, 1994; pilskaln et al., 1998), has been hypothesized to rival storms as the main agent for sediment resuspension and transport on the middle and outer continental shelf of the middle atlantic bight (churchill, 1989). whether indeed sediment resuspension induced by bottom trawling has more or less relevance than natural resuspension (as in the case of storms) in fuelling the water column with regenerated nutrients and suspended organic particles available as food for suspension feeders, remains to date a still largely unexplored issue. to provide insights on this issue, we investigated concentrations and biochemical composition of both total and bioavailable (i.e., enzymatically hydrolizable) pools of suspended organic particles in the thermaikos gulf (mediterranean sea) along a putative decreasing gradient of anthropogenic influence, during three periods: september 2001 (calm sea conditions and no trawling), october 2001 (trawling period), and february 2002 (no trawling, after a severe storm). more in details, we tested the null hypothesis by which the quantity, biochemical composition and bioavailability of suspended particles along the whole water column are not affected by bottom trawling or severe storms at the sea surface. methods study area and sampling the thermaikos gulf (fig. 1) is a micro-tidal environment, located in the north western aegean sea (eastern mediterranean), from 39°30’n to 40°38’n and 22°30’e to 23°19’e, with depths ranging from 30 to 200 m. the main circulation is characterized by more saline waters entering the outer shelf of the gulf over the eastern part, then turning towards the northeast in the inner part; less saline waters flow southerly along the western coastline (poulos et al., 2000). the thermaikos gulf is characterized by an extended shelf (180 km long and 55 km wide) of very smooth relief and by meso-eutrophic conditions (zervakis et al., 2005). the area under scrutiny, in particular, is characterized in its inner part by strong anthropogenic influences, due to the thessaloniki city’s harbour and the adjacent industrial zone. moreover, the gulf receives important riverine inputs from three major rivers (axios, aliakmon and pinios rivers; karageorgis and anagnostou, 2001). five sampling stations (namely ip01, ip10, ip17, ip38, ip41 at 30, 41, 55, 51, and 54 m depth, respectively) were located along the coast and positioned along a northto-south transect characterised by seasonally intensive bottom trawling activities (fig. 1). at each station, water samples were collected at 2 m, 20 m and about 1 m above the bottom, in september, october 2001 and in february 2002. these periods were selected as putatively representative of calm conditions and no trawling (september 2001), calm conditions with trawling (october 2001) and no trawling after a prolonged period of severe storms (february 2002). water samples were collected by means of a rosette sampler equipped with 15 l niskin bottles. after pre-filtration through a 200 µm mesh net to remove larger zooplankton, the water samples were filtered onto whatman gf/f filters (pre-combusted at 450°c, 4 h), immediately after collection. filters were stored at -20°c, until analyses. data obtained from the bottom layer of the water colfig. 1. study area and location of the sampling stations. no nco mm er cia l u se on ly 23 a. pusceddu et al. umn have been already partially discussed elsewhere (pusceddu et al., 2005b). in this study we extended the analysis including also the data obtained from the intermediate and surface layers of the water column, to document, if any, the effects of natural and anthropogenic sediment resuspension on the quantity and food availability of (re)suspended organic particles along the entire water column. biochemical composition of particulate organic matter concentrations of total particulate proteins (tprt), total particulate carbohydrates (tcho) and total particulate lipids (tlip) were determined according to hartree (1972) modified by rice (1982) (proteins), dubois et al. (1956) (carbohydrates), and marsh and weinstein (1966) and bligh and dyer (1959) (lipids), respectively. concentrations of the hydrolysable fractions of particulate proteins (hprt) and carbohydrates (hcho) were measured according to dell’anno et al. (2000), slightly modified for particulate samples (pusceddu et al., 2005c). filters were homogenised in 0.1 m na-phosphate buffer (ph 7.5), sonicated three times for 1 min (with intervals of 30 s) before enzyme addition. duplicate filters from each water depth (i.e., treated samples) were added to 100 µl of proteinasek (1 mg ml–1) and 100 µl of protease (600 µg ml–1). an equal volume of na-p buffer solution, without enzymes (i.e., control samples), was added to another set of filters. all the samples were incubated for 1 h at 37°c, under gentle agitation; they were filtered subsequently onto gf/f filters and rinsed twice with 5 ml of cold 0.1 m na-p buffer (ph 7.5), in order to remove the digested protein fraction and the surplus of enzymes. filters muffled at 450°c for 4 h and processed as described above were utilised as blanks. hydrolysable protein analyses from these samples were carried out spectrophotometrically as described above. difference in protein concentration between the control and treated samples were assumed to represent the concentration of proteins actually hydrolysed by proteases (hydrolysed proteins, hprt). for the enzymatic digestion of particulate carbohydrates, the filters were homogenised with 0.1 m na-phosphate, 0.1 m edta (ph 5.0) and sonicated three times for 1 min (with intervals of 30 s). replicate filters (n=3, treated samples) were added to 100 µl of α-amylase, 50 µl of β-glucosidase, 100 µl of proteinase-k and 100 µl of lipase (stock solution of all enzymes was 1 mg ml–1). another set of filters were treated by adding 0.1m na phosphate, instead of enzyme solutions, and utilised as a control. filters muffled at 450°c for 4 h and processed as described above, were utilised as blanks. after incubation, all the samples were centrifuged at 2000 g for 10 min and an aliquot of the supernatant was used to determine carbohydrates released from pom hydrolysis. soluble carbohydrates were determined from the supernatant of the control sample. carbohydrates from all the supernatants were analysed spectrophotometrically, as described above. the actual fraction of enzymatically hydrolysed carbohydrates (hcho) was obtained on the basis of the difference between the carbohydrate concentrations determined in the supernatant of samples containing enzymes and the soluble fraction of the control. total and hydrolysable carbohydrate and protein and total lipid concentrations were converted into carbon equivalents, using 0.40, 0.49 and 0.75 mgc mg–1 conversion factors, respectively (fabiano et al., 1995). the sum of the total protein, carbohydrate and lipid carbon equivalents was reported as particulate biopolymeric organic carbon (bpc, fabiano and pusceddu, 1998). particulate bioavailable organic carbon (baoc), as a proxy of the organic carbon potentially available for consumers (pusceddu et al., 2003, 2009), was defined as the sum of carbon equivalents of hydrolysable carbohydrates and proteins (danovaro et al., 2001). statistical analyses to test the null hypothesis that the quantity, biochemical composition and bioavailability of suspended particles along the whole water column are not affected by bottom trawling nor by severe storms at the sea surface, three-way permutational analyses of variance (permanova; anderson, 2001; mcardle and anderson, 2001) were carried out separately for each variable. the design included three orthogonal factors: period (p, 3 fixed levels: calm, trawling, storm), station (s, 5 fixed levels: ip01, ip10, ip17, ip38, ip41), and water depth (d, 3 fixed levels: surface, intermediate, bottom), with n=3 for the combination of factors. in the multivariate context, variations in the biochemical composition of suspended particles were assessed separately for the three water column layers using a 2-way design with period (p, 3 fixed levels: calm, trawling, storm), and station (s, 5 fixed levels: ip01, ip10, ip17, ip38, ip41) as orthogonal sources of variance. the analyses were based on euclidean distances of previously normalized data, using 4999 random permutations of the appropriate units (anderson and ter braak, 2003). because of the restricted number of unique permutations in the pairwise tests, p values were obtained from monte carlo samplings (anderson and robinson, 2003). significant differences among water layers in each station and sampling period, as well as differences among sampling periods in each station and water column layer were then assessed using snk post-hoc tests. canonical analysis of principal coordinates (cap) was used in the multivariate context to ascertain the allocation of experimental groups to those established a priori. results from the cap were then used to visualize, using biplots, differences among experimental groups (i.e., among periods and stations). the permanova and cap analyses were performed using the routines included in the primer 6+ software (clarke and gorley, 2006). no nco mm er cia l u se on ly 24bottom trawling impacts on particulate organic matter results total protein, carbohydrate, lipid, biopolymeric c, hydrolysable protein and carbohydrate and bioavailable organic carbon concentrations at each stations, water depth and sampling period are reported in tab. 1. the results of the permanova tests revealed a significant interaction of the three tested factors for all of the investigated variables (tab. 2). therefore, in order to identify the eventual significance of the major factor under scrutiny (i.e., trawling vs storm effects), we used post-hoc snk tests to discriminate differences in the concentration of suspended organic matter concentrations: i) among sampling depths during each period and at each station; ii) among sampling periods at each station and water column layer. the results of the post-hoc tests carried out to identify changes in the vertical distribution of suspended particles in the water column during the three different conditions (tab. 3) reveal: i) the presence during calm conditions in september of a nepheloid layer (i.e., significantly higher concentrations in the bottom layer of the water column) in almost all stations; ii) a more homogeneous distribution of pom (i.e., values in intermediate and surface layers higher than or similar to those in the bottom layer) during trawling activities in october and, though to a lesser extent, after-storm in february. these trends apply to almost all investigated variables and appear to be particularly evident at the shallowest stations ip01 and ip10. the post-hoc tests carried out to identify variations in the concentration of suspended organic compounds among sampling periods in each water column layer and in all sampling stations (tab. 4) reveal that: i) the stronger effects of bottom trawling when compared to calm and after-storm conditions are generally most evident in the intermediate and bottom layers of the water column, especially for the carbohydrate pools; ii) in some stations (i.e., ip10, ip17, ip38) and pre-eminently for the protein pools, suspended particle concentrations during trawling in october are similar to those observed after-storm in february and consistently higher than those during calm conditions in september. the results of the 2-way permanova conducted to identify variations in the biochemical composition of particulate organic matter among calm, trawling and afterstorm conditions and among sampling stations in each of the three layers of the water column reveal the presence of a significant period × station interaction for each water column layer (tab. 5). the bi-plots produced after the cap analysis to better ascertain, separately for each layer of the water column, changes in the biochemical composition of particulate organic matter during the three sampling periods reveal: i) in the surface layer of the water column a clear segregation of trawling conditions at all stations, with exception of ip17 and ip38, mostly driven by increasing concentrations of carbohydrate pools, whereas storm conditions mostly overlap with calm conditions (fig. 2a); ii) in the intermediate and bottom layers of the water column a general segregation of trawling and storm (slightly overlapped one each other) from calm conditions, mostly explained by decreasing concentrations of total particulate lipids and increasing concentrations of total and hydrolisable carbohydrate and protein pools during trawling and after storm (fig. 2b-c); iii) in the bottom layer of the water column, the best segregation among trawling, after-storm and calm conditions in the deepest stations (i.e., ip38 and ip41) (fig. 2c). in the surface layer of the water column of all stations, with exception of ip17 and ip38, the bioavailable fraction of particulate biopolymeric c is much higher during trawling activities than in calm and after-storm conditions (fig. 3a). the positive effect of trawling activities on the bioavailability of particulate biopolymeric c observed in the surface layer of the water column is smoother in the intermediate layer of the water column, where it is evident only in the deepest station ip41 (fig. 3b). a higher bioavailability of particulate biopolymeric c during trawling activities when compared to calm and after-storm conditions is again evident in the bottom layer of the water column, in the deepest stations ip38 and ip41 (fig. 3c). discussion we show here that sediment resuspension caused by intensive bottom trawling can determine effects on particulate organic matter (pom) quantity, biochemical composition and bioavailability which are similar or even stronger than those eventually exerted by storms at the sea surface. on the one hand, this result is in accordance with previous findings from the middle atlantic bight, which postulated that bottom trawling can rival natural resuspension induced by storm conditions (churchill, 1989). previous studies, conducted in the thermaikos gulf and, comparatively, in the gulf of lions (nw mediterranean sea), reported changes in the quantitative characteristics of sinking pom, resulting in increased total suspended matter concentrations and gross sedimentation rates through alternate cycles of resuspension and sedimentation associated with natural temporal variability (grémare et al., 2003; karageorgis and anagnostou, 2001; fernandes et al., 2009; zeri et al., 2009). our results show also that the bottom layer of the water column in the thermaikos gulf during calm conditions is characterised generally by concentrations of pom (and almost all its biochemical constituents) higher than those in the intermediate and upper layers. the presence of such a nepheloid layer indicates that natural background levels of sediment resuspension in the thermaikos gulf could be relatively high. in this sense, we must thereno nco mm er cia l u se on ly 25 a. pusceddu et al. ta b. 1 .t ot al a nd h yd ro liz ab le p ro te in , t ot al a nd h yd ro liz ab le c ar bo hy dr at e, to ta l l ip id , b io po ly m er ic c (b pc ), an d bi oa va ila bl e c c on ce nt ra tio ns a t a ) s ur fa ce (2 m ), b ) i nt er m ed ia te (2 0 m ) a nd c ) b ot to m d ep th s in th e t he rm ai ko s g ul f u nd er c al m (s ep te m be r 2 00 1) , d ur in g tr aw lin g (o ct ob er 2 00 1) a nd a ft er s to rm (f eb ru ar y 20 02 ) c on di tio ns . p ro te in c ar bo hy dr at e to ta l l ip id b io po ly m er ic c b io av ai la bl e c s ta ti on c on di ti on to ta l h yd ro liz ab le to ta l h yd ro liz ab le µ g l –1 sd µg l –1 sd µ g l –1 sd µg l –1 sd µ g l –1 sd µ gc l –1 s d µ gc l –1 s d % o f b p c a ) i p0 1 c al m 8 3. 7 1 .3 2 0. 4 1 .5 6 7. 7 1 .9 1 4. 7 0 .1 2 3. 1 0 .0 8 5. 4 1 .4 3 3. 2 0 .8 3 8. 8 t r aw lin g 1 00 .2 0 .5 5 1. 3 4 .0 8 0. 0 5 .4 2 8. 2 0 .3 2 3. 2 3 .3 9 8. 5 4 .9 5 3. 8 4 .5 5 4. 6 st or m 1 17 .8 7 .5 8. 8 0 .1 10 6. 7 3. 2 8 .5 2. 1 2 3. 2 3 .3 11 7. 8 7. 5 2 5. 1 3 .4 2 1. 3 ip 10 c al m 3 8. 1 2 .1 2. 1 0 .3 3 2. 3 1 .4 0. 0 0 .0 1 0. 1 0 .0 3 9. 1 1 .6 8 .6 0. 2 2 1. 9 tr aw lin g 99 .6 8 .4 6 6. 6 12 .2 1 26 .2 16 .3 3 8. 4 2 .0 9 .4 4. 7 10 3. 8 1 5. 7 52 .5 1 1. 8 50 .6 st or m 1 08 .5 6 .6 3. 8 0 .1 10 7. 1 4. 9 1 .0 0. 0 9 .3 4. 7 10 0. 5 1 0. 2 6 .8 5. 1 6 .7 ip 17 c al m 5 1. 0 0 .3 1 1. 2 1 .7 2 9. 1 1 .6 6. 5 1 .7 1 4. 3 2 .4 4 7. 3 2 .6 1 8. 8 3 .3 3 9. 7 tr aw lin g 59 .5 2 .0 0. 6 0 .0 6 7. 0 10 .8 0 .0 0. 0 7 .2 1. 4 6 1. 3 6 .3 5 .7 1. 0 9 .3 st or m 80 .4 4 .4 1. 0 0 .0 9 5. 9 4 .5 6. 9 2 .4 7 .2 1. 4 8 3. 1 4 .9 8 .6 2. 0 1 0. 4 ip 38 c al m 5 5. 5 0 .3 9. 4 0 .4 4 6. 9 7 .4 3 3. 1 0 .3 1 1. 3 2 .4 5 4. 4 4 .9 2 6. 3 2 .1 4 8. 3 tr aw lin g 65 .9 2 .3 1 0. 0 0 .6 8 0. 9 22 .4 1 1. 6 0 .0 7 .2 0. 0 7 0. 0 10 .1 14 .9 0 .3 2 1. 3 st or m 1 32 .1 8 .7 8 3. 3 4 .7 4 8. 4 2 .1 1. 0 0 .0 3 .1 1. 7 8 6. 4 6 .4 4 3. 5 3 .6 5 0. 4 ip 41 c al m 5 7. 6 6 .2 1 6. 0 0 .5 3 8. 6 2 .2 4. 9 0 .8 1 4. 1 2 .1 5 4. 3 5 .5 2 0. 3 2 .1 3 7. 5 tr aw lin g 76 .1 1 6. 8 23 .2 4 .1 9 8. 8 23 .4 5 4. 8 0 .4 3 .1 1. 7 7 9. 1 18 .9 35 .6 3 .5 4 5. 0 st or m 84 .0 1 8. 2 26 .5 2 .2 7 7. 2 10 .4 1 2. 3 3 .3 5 .0 0. 0 7 5. 8 13 .1 21 .6 2 .4 2 8. 5 b ) i p0 1 c al m 12 8. 4 3 1. 9 46 .8 0 .8 7 1. 3 2 .4 5. 2 0 .1 3 0. 9 2 .9 11 4. 7 1 8. 7 48 .2 2 .6 4 2. 1 tr aw lin g 1 32 .7 0 .5 4 5. 0 0 .2 9 1. 3 3 .1 0. 0 0 .0 2 1. 8 1 .4 11 7. 9 2. 5 3 8. 4 1 .2 3 2. 6 st or m 1 26 .1 2 .3 1 9. 8 3 .2 9 7. 4 1 .4 1. 0 0 .0 2 1. 8 1 .4 11 7. 1 2. 8 2 6. 5 2 .7 2 2. 6 ip 10 c al m 4 8. 6 0 .7 1. 9 0 .3 3 6. 5 2 .8 0. 6 0 .0 8 .9 1. 0 4 5. 0 2 .2 7 .8 0. 9 1 7. 4 tr aw lin g 77 .7 1 .0 4 8. 5 3 .3 11 3. 6 1 8. 6 0 .0 0. 0 7 .3 1. 0 8 9. 0 8 .7 2 9. 3 2 .4 3 2. 9 st or m 84 .4 0 .9 1. 0 0 .0 7 4. 8 3 .2 1. 0 0 .0 7 .3 1. 0 7 6. 8 2 .5 6 .4 0. 7 8 .3 ip 17 c al m 3 7. 0 6 .4 0. 0 0 .0 2 7. 3 3 .6 5. 7 0 .6 1 3. 7 1 .0 3 9. 3 5 .3 1 2. 6 1 .0 3 2. 0 tr aw lin g 72 .3 0 .7 1 8. 7 0 .6 8 4. 8 2 .1 0. 0 0 .0 3 .6 2. 4 7 2. 1 3 .0 1 1. 9 2 .1 1 6. 5 st or m 70 .3 3 .3 7. 8 0 .4 9 3. 5 1 .2 3 2. 5 3 .5 3 .6 2. 4 7 4. 6 3 .9 1 9. 5 3 .4 2 6. 2 ip 38 c al m 8 1. 0 2 .8 1 8. 9 6 .6 4 8. 3 4 .1 1 1. 9 0 .7 1 5. 8 2 .0 7 0. 9 4 .5 2 5. 9 5 .1 3 6. 6 tr aw lin g 61 .4 5 .4 3. 2 0 .2 10 4. 1 8. 8 6 7. 8 1 .9 5 .0 1. 4 7 5. 5 7 .3 3 2. 4 1 .9 4 2. 9 st or m 1 05 .5 1 .3 4 7. 1 12 .7 65 .1 2 .2 2 1. 1 0 .1 3 .1 2. 4 8 0. 1 3 .3 3 3. 8 8 .0 4 2. 2 ip 41 c al m 11 3. 2 3. 7 0 .0 0. 0 4 9. 5 2 .9 0. 0 0 .0 1 6. 6 1 .0 8 7. 7 3 .7 1 2. 4 0 .8 1 4. 2 tr aw lin g 63 .8 4 .7 2 2. 2 5 .5 11 5. 0 9. 7 5 0. 6 14 .0 3 .1 2. 4 7 9. 6 8 .0 3 3. 4 10 .1 4 2. 0 st or m 70 .9 9 .1 2 3. 8 0 .3 9 3. 1 1 .3 1. 0 0 .0 2 .1 1. 2 7 3. 5 5 .9 1 3. 6 1 .0 1 8. 5 c ) i p0 1 c al m 17 3. 3 6. 1 7 6. 7 3 .7 15 0. 6 5. 8 7 3. 5 5 .3 4 0. 4 1 .0 17 5. 4 6. 0 9 7. 3 4 .7 5 5. 4 tr aw lin g 1 35 .1 10 .3 1 8. 8 1 .1 12 3. 4 1 7. 6 42 .0 0 .9 2 2. 2 1 .9 13 2. 2 1 3. 5 42 .6 2 .3 3 2. 2 st or m 1 85 .9 5 .2 4 6. 0 1 .9 15 1. 5 1 1. 0 67 .6 4 .1 2 2. 2 1 .9 16 8. 3 8. 4 6 6. 2 4 .0 3 9. 3 ip 10 c al m 8 4. 9 8 .0 1 2. 9 0 .7 8 5. 6 2 .2 3 9. 3 1 .9 2 9. 9 0 .5 9 8. 3 5 .2 4 4. 5 1 .4 4 5. 3 tr aw lin g 1 08 .2 1 .4 2 3. 8 5 .7 15 3. 0 3 6. 1 40 .2 9 .3 1 0. 0 2 .9 12 1. 7 1 7. 3 35 .3 8 .7 2 9. 0 st or m 1 05 .2 7 .5 2 8. 4 4 .9 10 5. 1 2. 5 1 .0 0. 0 1 0. 0 2 .9 10 1. 1 6. 8 2 1. 8 4 .5 2 1. 6 ip 17 c al m 7 6. 0 0 .9 2 1. 2 0 .7 8 4. 1 2 .5 4 8. 9 8 .2 2 7. 6 0 .0 9 1. 6 1 .5 5 0. 7 3 .7 5 5. 3 tr aw lin g 1 05 .9 6 .6 2 1. 1 2 .8 9 8. 1 14 .1 2 7. 3 6 .1 1 .3 0. 8 9 1. 7 9 .8 2 1. 8 4 .7 2 3. 8 st or m 1 35 .1 13 .1 5 0. 0 0 .8 14 9. 0 4. 2 1 .0 0. 0 4 .0 0. 4 12 8. 7 8. 4 2 7. 9 0 .7 2 1. 6 ip 38 c al m 10 3. 5 0. 5 2 6. 2 9 .6 7 4. 3 8 .0 4 4. 0 1 .0 1 7. 5 2 .9 9 3. 5 5 .6 4 3. 6 7 .2 4 6. 6 tr aw lin g 11 2. 8 4. 2 6 4. 0 6 .2 17 4. 5 4. 6 10 9. 8 1 9. 3 4 .0 0. 4 12 8. 1 4. 2 7 8. 2 11 .0 6 1. 0 st or m 1 28 .4 6 .6 5 3. 0 1 .3 11 4. 6 5. 6 7 9. 6 5 .4 6 .0 0. 0 11 3. 3 5. 5 6 2. 3 2 .8 5 5. 0 ip 41 c al m 8 8. 9 14 .6 1 .0 0. 1 6 5. 9 0 .2 2 5. 7 5 .5 3 4. 3 3 .8 9 5. 7 10 .1 36 .5 5 .1 3 8. 1 tr aw lin g 77 .3 4 .7 1 6. 1 1 .4 13 9. 0 7. 5 9 7. 0 7 .5 6 .0 0. 0 9 8. 0 5 .3 5 1. 2 3 .6 5 2. 2 st or m 1 00 .5 5 .6 3 9. 7 8 .0 5 7. 4 3 .4 7. 5 0 .6 6 .7 1. 9 7 7. 2 5 .5 2 7. 5 5 .6 3 5. 6 sd , s ta nd ar d de vi at io n (n =3 ). no nco mm er cia l u se on ly 26bottom trawling impacts on particulate organic matter tab. 2. results of 3-way permanova testing for differences in the investigated variables among sampling periods, stations and sampling depths. source df ms f p total protein period (p) 2 8.2 134.9 *** station (s) 4 12.1 199.4 *** depth (d) 2 14.1 232.4 *** p x s 8 1.5 24.2 *** p x d 4 1.3 21.7 *** s x d 8 1.2 19.1 *** p x s x d 16 0.6 9.9 *** residual 90 0.1 hydrolysable protein period (p) 2 4.0 117.6 *** station (s) 4 5.7 165.0 *** depth (d) 2 4.5 131.1 *** p x s 8 5.1 148.0 *** p x d 4 1.3 38.4 *** s x d 8 1.1 31.8 *** p x s x d 16 2.3 66.9 *** residual 90 0.0 total carbohydrate period (p) 2 21.4 292.7 *** station (s) 4 2.0 26.7 *** depth (d) 2 17.3 236.3 *** p x s 8 2.9 39.8 *** p x d 4 0.4 5.8 ** s x d 8 1.1 14.6 *** p x s x d 16 0.5 7.4 *** residual 90 0.1 hydrolysable carbohydrate period (p) 2 7.1 270.2 *** station (s) 4 4.5 173.7 *** depth (d) 2 19.7 755.9 *** p x s 8 2.9 109.4 *** p x d 4 0.7 25.2 *** s x d 8 1.7 66.0 *** p x s x d 16 1.3 49.6 *** residual 90 0.0 total lipid period (p) 2 21.2 397.8 *** station (s) 4 14.1 264.8 *** depth (d) 2 4.0 75.2 *** p x s 8 0.4 8.5 *** p x d 4 3.2 60.8 *** s x d 8 0.4 7.7 *** p x s x d 16 0.2 3.3 ** residual 90 0.1 biopolymeric c period (p) 2 4.9 66.8 *** station (s) 4 11.6 156.8 *** depth (d) 2 20.9 282.8 *** p x s 8 1.5 19.7 *** p x d 4 1.2 15.7 *** s x d 8 0.8 11.4 *** p x s x d 16 0.4 5.6 *** residual 90 0.1 bioavailable organic c period (p) 2 2.0 39.3 *** station (s) 4 9.3 183.6 *** depth (d) 2 19.6 387.0 *** p x s 8 2.4 47.8 *** p x d 4 1.1 21.5 *** s x d 8 0.6 12.8 *** p x s x d 16 1.3 25.2 *** residual 90 0.1 df, degrees of freedom; ms, mean square; f, f value; ***p<0.001; **p<0.01. fig. 2. bi-plots after cap analysis illustrating spatial (among stations) and temporal (among calm, trawling and after-storm conditions) variations in the biochemical composition of particulate organic matter in the thermaikos gulf, in the superficial (a), intermediate (b) and bottom (c) water layers. vectors are proportional to the correlation of variables with the two major axes. tprt, total proteins; tcho, total carbohydrates; tlip, total lipids; hprt, hydrolysable proteins; hcho, hydrolysable carbohydrates. no nco mm er cia l u se on ly 27 a. pusceddu et al. tab. 4. visual representation of post-hoc tests carried out to ascertain variations in the concentration of particulate organic compounds in the thermaikos gulf during calm (september, c), trawling (october 2001, t) and after storm (february 2002, s) conditions in the three water column layers. green/white, yellow, light blue and red cells indicate missing effects of both trawling and storm, stronger effects of storm, similar effects of storm and trawling, stronger effects of trawling, respectively. water layer variable ip1 ip10 ip17 ip38 ip41 surface total protein s>t>c s,t>c s,t>c s,t>c ns hydrolysable protein t>c>s t>s>c c>s>t s>c,t s,t>c total carbohydrate s>t>c s,t>c s>t>c ns s,t>c hydrolysable carbohydrate t>c>s t>s>c s>c>t c>t>s t>s>c total lipid ns ns c>t,s c>t>s c>t,s biopolymeric c s>t>c s,t>c s>t>c s>c ns bioavailable organic c t>c>s t>c>s c>s>t s>c>t t>c intermediate total protein t>s s>t>c s,t>c s>c>t c>t,s hydrolysable protein c>t>s t>c>s t>s>c s>c>t s,t>c total carbohydrate s>t>c t>s>c t>s>c t>s>c t>s>c hydrolysable carbohydrate c>s>t c>s>t s>c>t t>s>c t>s>c total lipid c>t,s ns c>t,s c>t,s c>t,s biopolymeric c ns s,t>c s,t>c s>c c>s bioavailable organic c c,t>s t>c>s s>c>t ns t>s>c bottom total protein c,s>t s,t>c t>s>c s>t>c s>t hydrolysable protein c>s>t s,t>c s>t,c s,t>c s>t>c total carbohydrate ns t>s>c s>t,c t>s>c t>s>c hydrolysable carbohydrate c,s>t c,t>s c>t>s s,t>c t>c>s total lipid c>t,s c>t,s c>s>t c>s>t c>s,t biopolymeric c c,s>t ns s>t,c t>s>c c,t>s bioavailable organic c c>s>t c>s c>t,s s,t>c t>c,s ns, not significant. tab. 3. visual representation of post-hoc tests carried out to ascertain variations in the distribution of particulate organic compounds in the water column of the thermaikos gulf during calm (september 2001), trawling (october 2001) and after storm (february 2002) conditions. green, yellow and red cells indicate missing, weak and strong signatures, respectively, of sediment resuspension in the upper water column layer. condition (period) variable ip01 ip10 ip17 ip38 ip41 calm (sept 2001) total protein b>s b>i>s b>s>i b>i>s b,i>s hydrolysable protein b>i>s b>i,s b>s>i b>s s>b>i total carbohydrate b>i,s b>i,s b>i,s b>i,s b>i>s hydrolysable carbohydrate b>s>i b>i>s b>i,s b>s>i b>s>i total lipid b>i>s b>i,s b>i,s b>s b>i,s biopolymeric c b>i,s b>i>s b>i,s b>i>s b,i>s bioavailable organic c b>i>s b>i,s b>s>i b>i,s b>s>i trawling (oct 2001) total protein b,i>s b,s>i b>i>s b>i,s b>i hydrolysable protein s>i>b s,i>b b,i>s b>s>i ns total carbohydrate b>i>s ns b>s b>i,s b>i hydrolysable carbohydrate b>s>i b,s>i b>s,i b>i>s b>i,s total lipid ns ns s>b s>b b>i,s biopolymeric c b,i>s b>i b>i,s b>i,s b>i bioavailable organic c s>b,i s>i b>i>s b>i>s b>s after storm (feb 2002) total protein b>s,i b,s>i b>s>i b,s>i b>i hydrolysable protein b>i>s b>s>i b>i>s s>i,b i>b total carbohydrate b>s>i b,s>i b>i,s b>i>s i>s>b hydrolysable carbohydrate b>s>i ns i>s>b b>i>s b,s>i total lipid ns ns s>b b>s b,s>i biopolymeric c b>s,i b,s>i b>s,i b>s,i ns bioavailable organic c b>s,i b>s,i b>i>s b>s,i b,s>i b, bottom; i, intermediate; s, surface. no nco mm er cia l u se on ly 28bottom trawling impacts on particulate organic matter fore acknowledge that a certain (unknown) proportion of variance among calm, trawling, and after-storm conditions is most likely associated to natural temporal variability in the pom quantity and biochemical composition in the thermaikos gulf. indeed, variations in the concentration of pom along the water column, as well as in its biochemical composition, can be the result of biological processes, including among the others primary productivity and particle consumption (fabiano and pusceddu, 1998; fabiano et al., 2001). moreover, concentration and composition of pom can be affected, in particular along the continental shelf, by rivers discharge (goñi et al., 2013). this latter, however, was most likely not the case as the analysis of meteorological data in the region during the study period revealed a very dry season in autumn 2001, accompanied by very low levels of river discharge (tragou et al., 2005) differences in the environmental characteristics among the sampling periods appeared to be particularly relevant between september-october 2001 and february 2002. in fact, the hydrological characteristics of the thermaikos gulf did not change markedly from september (calm conditions) to october 2001 (during trawling), when pronounced thermoclines and haloclines were recorded at ca. 35-40 m depth in the whole study area (tragou et al., 2005; zervakis et al., 2005). on the other hand, a series of cold fronts that passed over the region in late january 2002 completely homogenised the thermohaline structure of the water column in february 2002 (tragou et al., 2005; zervakis et al., 2005). moreover, during this period, a low salinity, low temperature water front occurred along the western boundary of the gulf, characterised by strong, vertically homogeneous southward velocities of up to 20 cm s–1 (zervakis et al., 2005). these large differences would make in principle difficult interpreting the differences in pom characteristics among the september-october 2001 and february 2002 periods. however, the lack of significant variations in the physicochemical characteristics from calm to trawling conditions, at least, allows us to corroborate the hypothesis by which bottom trawling is a major factor stimulating sediment resuspension and, as a consequence, increasing pom concentrations in the water column as well as significantly modifying its biochemical composition and bioavailability for consumers. this result is indeed consistent with previous investigations carried out by means of both correlative and manipulative approaches, which generally demonstrated that bottom trawling can significantly inject large amounts of sediments and associated pom into the overlying water column (durrieu de madron et al., 2005; pusceddu et al., 2005a, 2005c; martín et al., 2014b). our results paradigmatically demonstrate that along with the well-known mechanisms of pelagic-benthic coupling associated with sedimentation processes of pom (graf, 1992), the reverse exchange of material (i.e., benthicpelagic coupling, sensu marcus and boero, 1998) stimulated by sediment resuspension caused by anthropogenic and, to a lesser extent, natural processes can have important consequences on pom stocks in the water column. we report here also that the transition from calm to trawling and to after-storm conditions is characterised by clear changes in the biochemical composition of pom. in particular, we show that both during trawling in october and after-storm in february, the relative importance of total and hydrolysable carbohydrates increases significantly when compared to values measured during calm conditions. at the same time, suspended pom during trawling and after the storm is also characterised by lipid tab. 5. results of permanova analysis testing for differences in the biochemical composition of particulate organic matter considering all investigated variables in the three layers of the water column. source df ms f p surface period 2 34.302 85.383 *** station 4 15.271 38.012 *** period x station 8 9.7823 24.349 *** residual 30 0.40175 total 44 intermediate period 2 27.632 102.07 *** station 4 21.758 80.37 *** period x station 8 8.6978 32.128 *** residual 30 0.27072 total 44 bottom period 2 28.789 106.5 *** station 4 21.519 79.606 *** period x station 8 8.5294 31.553 *** residual 30 0.27032 total 44 df, degrees of freedom; ms, means square; f, f value; ***p<0.001. no nco mm er cia l u se on ly 29 a. pusceddu et al. concentrations significantly lower than those in calm conditions. the gross biochemical composition of pom is the result of a complex multiple-source array of biotic and abiotic factors (danovaro et al., 2000) and variations in the relative importance of protein, carbohydrate and lipid contents can be highly indicative of changes in the labile vs refractory nature of particles (fabiano et al., 2001). carbohydrates are usually associated with organic matter pre-eminently refractory in nature (grémare et al., 2003; pusceddu et al., 2009), so that our results strongly suggests that sediment resuspension caused by trawling and, though to a lesser extent, by storms can lower the food availability of suspended organic particles. this hypothesis is also corroborated by the significant decrease, during trawling, in the concentration of particulate lipids, whose major fraction is generally highly labile (carreira et al., 2010). altogether these results would suggest that om particles resuspended after bottom trawling could more refractory in nature than those present in the water column in calm or after-storm conditions. nevertheless, we report here that, in contrast with what hypothesized from the gross biochemical composition only, the bioavailable fraction of particulate biopolymeric c increases significantly during trawling, though not consistently at all stations and water layer. this apparent incongruence is due to the fact that while total protein, lipid and carbohydrate pools include generally a very heterogeneous complex of organic compounds which are not fig. 3. variations in the bioavailable fraction of particulate organic matter (in terms of percentage fraction of biopolymeric c enzymatically digestible) in the thermaikos gulf during calm (september 2001), trawling (october 2001) and after-storm (february 2002) conditions. no nco mm er cia l u se on ly 30bottom trawling impacts on particulate organic matter equally reactive to degradation (pusceddu et al., 2009), the enzymatically digestible fractions of protein and carbohydrate pools (i.e., the bioavailable fraction of organic c; dell’anno et al., 2000; pusceddu et al., 2003) are more reliable descriptors of organic matter food availability to consumers than the gross biochemical composition alone (fabiano and pusceddu, 1998; pusceddu et al., 1999; danovaro et al., 2001). the apparent positive effect of bottom trawling on the bioavailability of suspended pom is however, most clearly confined to the bottom layer of the deepest stations investigated in this study. this result does not allow us concluding that sediment resuspension induced by trawling can cause the injection of labile organic compounds in the whole water column, nor that the faster mobilisation rates of organic c buried in the sediment after trawling in coastal sediments is a major factor contributing to coastal eutrophication (polymenakou et al,. 2005; pusceddu et al., 2005c). previous studies carried out in the field and under laboratory conditions have demonstrated that sediment resuspension caused by storms or trawling activities can enhance either suspended or sedimentary pom quantity and bioavailability (pusceddu et al., 2005a, 2005b, 2005c). in particular, benthic microbes exposed to o2-rich waters caused by sediment disturbance induced by trawling can stimulate a faster mobilisation of organic c buried in the sediment, injecting more labile molecules into the system (polymenakou et al., 2005). other studies have also demonstrated that bottom trawling in coastal waters creates plumes of resuspended sediments which last just a few hours in the water column (schoellhamer, 1996; palanques et al., 2001; durrieu de madron et al., 2005). nevertheless, volumes moved from the sediment to the water column can be huge if trawling activities are chronic as observed in several regions of the mediterranean sea (martín et al., 2014a). in such conditions we can hypothesize that the overall enhancement of c cycling in the water column and in the upper layers of the sediment exposed to chronic bottom trawling might stimulate an increase of available nutrients able, in turn, to sustain increased levels of primary productivity, ultimately leading to a potential internal eutrophication process (polymenoakou et al., 2005). overall, the results of this study are very different from what, instead, observed in deep-sea sediments exposed to chronic trawling activities, where the lack of a conspicuous re-deposition from the water column determines a dramatic lowering of organic c sedimentary contents and turnover rates (pusceddu et al., 2014). conclusions although limited to a very short-term analysis, likely biased by the uncontrolled seasonal variations in the quantity and biochemical composition of pom, our results corroborate the most recent literature which demonstrates that bottom trawling represents a major threat not only for commercially exploited target species and the associated by-catch, but can have also important consequences on the biogeochemical cycling of organic c. our results have also shown that the effects of intensive trawling activities can rival those (temporarily) exerted by natural events (e.g., storms). bottom trawling represents the most common fishing practice worldwide, it is being carried out at progressively deeper depths (puig et al., 2012) and is exerting severe consequences on the submersed seascape (martín et al., 2014a) as well on the biodiversity and functioning of marine ecosystems (pusceddu et al., 2014). since bottom trawling is carried out worldwide and natural storms at sea can be frequent and intense, we claim for the need of assessing new adapting management strategies of bottom trawling in order to mitigate the synergistic impacts of anthropogenic and natural sediment resuspension on coastal biogeochemical cycles.. acknowledgments this work has been supported by the project interpol (evk3-2000–0023) of the european commission. references althaus f, williams a, schlacher ta, kloser rj, green ma, barker ba, bax nj, brodie p, schlacher-hoenlinger ma, 2009. impacts of bottom trawling on deep-coral ecosystems of seamounts are long-lasting. mar. ecol. prog. ser. 397: 279-294. anderson mj, 2001. a new method for non-parametric multivariate analysis of variance. austral. ecol. 26:32-46. anderson mj, robinson j, 2003. generalised discriminant analysis based on distances. aust. n.z. j. stat. 45:301-318. anderson mj, ter braak cjf, 2003. permutation tests for multifactorial analysis of variance. j. stat. comput. sim. 73: 85-113. black kp, parry gd, 1994. sediment transport rates and sediment disturbance due to scallop dredging in port phillip bay. mem. queensl. mus. 36:327-341. bligh eg, dyer wj, 1959. a rapid method for total lipid extraction and purification. can. j. biochem. physiol. 37:911-917. bongiorni l, mea m, gambi c, pusceddu a, taviani m, danovaro r, 2010. deep-water scleractinian corals promote higher biodiversity in deep-sea meiofaunal assemblages along continental margins. biol. conserv. 143:1687-1700. bruckner aw, 2009. rate and extent of decline in corallium (pink and red coral) populations: existing data meet the requirements for a cites appendix ii listing. mar. ecol. progr. ser. 397:319-332. carreira rs, araújo mp, costa tlf, ansari nr, pires lcm, 2010. lipid biomarkers in deep sea sediments from the campos basin, se brazilian continental margin. org. geochem. 41:879-884. no nco mm er cia l u se on ly 31 a. pusceddu et al. christensen jp, 1989. sulfate reduction and carbon oxidation rates in continental shelf sediments, an examination of offshelf carbon transport. cont. shelf res. 9:223-246. churchill jh, 1989. the effect of commercial trawling on sediment resuspension and transport over the middle atlantic bight continental shelf. cont. shelf res. 9:841-864. clarke kr, gorley rn, 2006. primer v6: user manual/tutorial. primer-e, plymouth, p.190. danovaro r, dell’anno a, pusceddu a, marrale d, della croce n, fabiano m, tselepides a, 2000. biochemical composition of pico-, nanoand micro-particulate organic matter and bacterioplankton biomass in the oligotrophic cretan sea (ne mediterranean). progr. oceanogr. 46:279-310. danovaro r, dell’anno a, fabiano m, 2001. bioavailability of organic matter in the sediments of the porcupine abyssal plain. mar. ecol. progr. ser. 220:25-32. dell’anno a, fabiano m, mei ml, danovaro r, 2000. enzymatically hydrolysed protein and carbohydrate pools in deep-sea sediments: estimates of the potential bioavailable fraction and methodological considerations. mar. ecol. progr. ser. 196:15-23. dubois m, gilles ka, hamilton jk, rebers pa, smith f, 1956. colorimetric method for determination of sugars and related substances. anal. chem. 28:350-356. durrieu de madron x, ferre b, le corre g, grenz c, conan p, pujo pay m, buscail r, bodiot o, 2005. trawling-induced resuspension and dispersal of muddy sediments and dissolved elements in the gulf of lion (nw mediterranean). cont. shelf res. 25:2387-2409. fabiano m, danovaro r, fraschetti s, 1995. a three-year time series of elemental and biochemical composition of organic matter in subtidal sediments of the ligurian sea (northwestern mediterranean). cont. shelf res. 15:1453-1469. fabiano m, pusceddu a, 1998. total and hydrolizable particulate organic matter (carbohydrates, proteins and lipids) at a coastal station in terra nova bay (ross sea, antarctica). polar biol. 19:125-132. fabiano m, pusceddu a, dell’anno a, armeni m, vanucci s, lampitt r, wolff g, danovaro r, 2001. fluxes of phytopigments and labile organic matter to the deep ocean in the ne atlantic ocean. progr. oceanogr. 50:89-104. fanning ka, carder kl, betzer pr, 1982. sediment resuspension by coastal waters: a potential mechanism for nutrient re-cycling on the ocean’s margins. deep sea res. 29:953-965. fernandes l, bhosle nb, prabhu matondkar sg, bhushan r, 2009. seasonal and spatial distribution of particulate organic matter in the bay of bengal. j. mar. sys. 77:137-147. fowler cw, 2003. tenets, principles, and criteria for management: the basis for systemic management. mar. fish. rev. 65:1-55. goñi ma, hatten ja, wheatcroft ra, borgeld jc, 2013. particulate organic matter export by two contrasting small mountainous rivers from the pacific northwest, u.s.a. j. geophys. res. biogeosci. 118:112-134. graf g, 1992. benthic-pelagic coupling: a benthic view. oceanogr. mar. biol. ann. rev. 30:149-190. grémare a, amouroux jm, cauwet g, charles f, courties c, de bovée f, dinet a, devenon jl, durrieu de madron x, ferre b, fraunie p, joux f, lantoine f, lebaron p, naudin jj, palanques a, pujo-pay m, zudaire l, 2003. the effects of a strong winter storm on physical and biological variables at a shelf site in the mediterranean. oceanol. acta 26:407-419. hartree ef, 1972. determination of proteins: a modification of the lowry method that gives a linear photometric response. anal. biochem. 48:422-427. heifetz j, stone rp, shotwell sk, 2009. damage and disturbance to coral and sponge habitat of the aleutian archipelago. mar. ecol. progr. ser. 397:295-303. hiddink jg, jennings s, kaiser mj, 2007. assessing and predicting the relative ecological impacts of disturbance on habitats with different sensitivities. j. appl. ecol. 44:405-413. hinz h, prieto v, kaiser mj, 2009. trawl disturbance on benthic communities: chronic effects and experimental predictions. ecol. appl. 19:761-773. jennings s, pinnegar jk, polunin nvc, warr kj, 2001. impacts of trawling disturbance on the trophic structure of benthic invertebrate communities. mar. ecol. progr. ser. 213:127-142. johnson af, gorelli g, jenkins sr, hiddink jg, hinz h, 2015. effects of bottom trawling on fish foraging and feeding. proc. r. soc. b 282:20142336. kaiser m, 1998. significance of bottom fishing disturbance. conserv. biol. 12:1230-1235. karageorgis ap, anagnostou cl, 2001. particulate matter spatial-temporal distribution and associated surface sediment properties: thermaikos gulf and sporades basin, nw aegean sea. cont. shelf res. 21:2141-2153. marcus nh, boero f, 1998. minireview: the importance of benthic-pelagic coupling and the forgotten role of life cycles in coastal aquatic systems. limnol. oceanogr. 43:763-768. marsh jb, weinstein db, 1966. a simple charring method for determination of lipids. j. lipid res. 7:574-576. martín j, puig p, palanques a, giamportone a, 2014a. commercial bottom trawling as a driver of sediment dynamics and deep seascape evolution in the anthropocene. anthropocene 7:1-15. martín j, puig p, palanques a, ribó m, 2014b. trawling-induced daily sediment resuspension in the flank of a mediterranean submarine canyon. deep sea res. ii 104:174-183. mcardle bh, anderson mj, 2001. fitting multivariate models to community data: a comment on distance-based redundancy analysis. ecology 82:290-297. palanques a, guillén j, puig p, 2001. impact of bottom trawling on water turbidity and muddy sediment of an unfished continental shelf. limnol. oceanogr. 46:1100-1110. pilskaln ch, churchill jh, mayer lm, 1998. resuspension of sediment by bottom trawling in the gulf of maine and potential geochemical consequences. conserv. biol. 12:1223-1229. polymenakou pn, pusceddu a, tselepides a, polychronaki t, giannakourou a, fiordelmondo c, hatziyanni e, danovaro r, 2005. benthic microbial abundance and activities in an intensively trawled ecosystem (thermaikos gulf, aegean sea). cont. shelf res. 25:2570-2584. poulos se, chronis gth, collins mb, lykousis v, 2000. thermaikos gulf coastal system, nw aegean sea: an overview of water/sediment fluxes in relation to air-land-ocean interactions and human activities. j. mar. syst. 25:47-76. puig p, canals m, company jb, martín j, amblas d, lastras g, palanques a, calafat am, 2012. ploughing the deep sea floor. nature 489:286-289. pusceddu a, bianchelli s, martín j, puig p, palanques a, masqué p, danovaro r, 2014. chronic and intensive bottom no nco mm er cia l u se on ly 32bottom trawling impacts on particulate organic matter trawling impairs deep-sea biodiversity and ecosystem functioning. proc. natl. acad. sci. 111:8861-8866. pusceddu a, dell’anno a, fabiano m, danovaro r, 2009. quantity and bioavailability of sediment organic matter as signatures of benthic trophic status. mar. ecol. progr. ser. 375:41-52. pusceddu a, dell’anno a, manini e, fabiano m, sarà g, danovaro r, 2003. enzymatically hydrolysable protein and carbohydrate sedimentary pools as indicators of the trophic state of ‘detritus sink’ systems: a case study in a mediterranean coastal lagoon. estuaries 26:641-650. pusceddu a, fiordelmondo c, danovaro r, 2005a. sediment resuspension effects on the benthic microbial loop in experimental microcosms. microb. ecol. 50:602-613. pusceddu a, fiordelmondo c, polymenakou p, polychronaki t, tselepides a, danovaro r, 2005b. effects of bottom trawling on the quantity and biochemical composition of organic matter in coastal marine sediments (thermaikos gulf, northwestern aegean sea). cont. shelf res. 25:2491-2505. pusceddu a, grémare a, escoubeyrou k, amouroux jm, fiordelmondo c, danovaro r, 2005c. impact of natural (storm) and anthropogenic (trawling) sediment resuspension on particulate organic matter in coastal environments. cont. shelf res. 25:2506-2520. pusceddu a, sarà g, armeni m, fabiano m, mazzola a, 1999. seas on a land spatial changes in the sediment organic matter of a semi-enclosed marine system (w-mediterranean sea). hydrobiologia 397:59-70. queiros am, hiddink jg, kaiser mj, hinz h, 2006. effects of chronic bottom trawling disturbance on benthic biomass, production and size spectra in different habitats. j. exp. mar. biol. ecol. 335:91-103. rice dl, 1982. the detritus nitrogen problem: new observations and perspectives from organic geochemistry. mar. ecol. progr. ser. 9:153-162. riemann b, hoffmann e, 1991. ecological consequences of dredging and bottom trawling in the limfjord, denmark. mar. ecol. progr. ser. 69:171-178. schoellhamer dh, 1996. anthropogenic sediment resuspension mechanisms in a shallow microtidal estuary. estuar. coas. shelf sci. 43:533-548. smith be, collie js, lengyel nl, 2013. effects of chronic bottom fishing on the benthic epifauna and diets of demersal fishes on northern georges bank. mar. ecol. prog. ser. 472:199-217. smith cj, rumohr h, karakassis i, papadopoulou kn, 2003. analysing the impact of bottom trawls on sedimentary seabeds with sediment profile imagery. j. exp. mar. biol. ecol. 285-286: 479-496. thrush sf, dayton pk, 2002. disturbance to marine benthic habitats by trawling and dredging: implication for marine biodiversity. annu. rev. ecol. syst. 33:449-473. tragou e, zervakis v, papageorgiou e, stavrakakis s, lykousis v, 2005. monitoring the physical forcing of resuspension events in the thermaikos gulf-nw aegean during 20012003. cont. shelf res. 25:2315-2331. zeri c, kontoyiannis h, giannakourou a, 2009. distribution, fluxes and bacterial consumption of total organic carbon in a populated mediterranean gulf. cont. shelf res. 29:886-895. zervakis v, karageorgis ap, kontoyiannis h, papadopoulos v, lykousis v, 2005. hydrology, circulation and distribution of particulate matter in thermaikos gulf (nw aegean sea), during september 2001-october 2001 and february 2002. cont. shelf res. 25:2332-2349. no nco mm er cia l u se on ly layout 1 introduction most typically, the environment is spatially structured in relation to different and variable energy inputs, which altogether modulate the distribution of organisms in patches or gradients (legendre and fortin, 1989; levin and sibuet 2012; woolley et al., 2016). indeed, the typical non-random pattern of species or species assemblages’ organization determines the aggregation into clumps or patches of abundance (levin and sibuet, 2012; pinckney and sandulli, 1990; rex and etter, 2010). the patchy distribution of marine benthos is the result of many different factors, including, among the others, either abiotic (e.g., geomorphology, hydrodynamic conditions; zeppilli et al., 2016) or biotic (e.g., resource availability, competition and predation; denny et al., 2004) factors. the knowledge about the complexity of natural habitats is essential to understand the characteristics of assemblages’ patchy distribution (legendre and fortin, 1989). the heterogeneity of the physical environment contributes significantly to preserve the community diversity, as well as the diversity of biological processes and ecological interactions that can be observed at different points in space (legendre and fortin, 1989). for instance, in marine ecosystems, habitat spatial heterogeneity and discontinuity of the substrates can explain the heterogeneous distribution of meiofauna and the relationships between biodiversity and ecosystem functioning (zeppilli et al., 2016). the comparison between spatial models of distribution of consumers and their resources gives information about the trophic interactions and the scale at which these interactions occur (sandulli and pickney, 1999). the scales of the spatial variability of the marine benthos have been investigated since a long time. however, while the scales of spatial variability for macrobenthos are well known, those of meiofauna and prokaryotes are still matter of debate (danovaro et al., 2001; fontaneto and hortal, 2012; moens et al., 2013; pusceddu et al., 2014; prat et al., 2015). meiofauna show frequently an irregular distribution over different spatial scale (moens et al., 2013; cerca et al., 2018; fontaneto, 2019;) several causes have been indicated to explain this, in particular for microscale variations (i.e., centimeters), and include, among the others: hydrodynamics (semprucci et al., 2011), biotic structures (bianchelli et al., 2013), reproduction, predation, micro-topography (raes et al., 2007 ), intraspecific interactions (chandler and flegger, 1987) and food resources (semprucci et al., 2019). one of the potential food sources for meiofauna is represented by the sedimentary organic matter (om), which also includes the fraction originating from the micro-algae. om quantity and biochemical composition in marine sediments allows to determine the trophic state of the system (dell’anno et al., 2002; pusceddu et al., 2004, 2009), i.e. the levels of potential food availability for the benthic consumers and, therefore, for the meiofauna (pusceddu et al., 2007, 2011). article small-scale distribution of metazoan meiofauna and sedimentary organic matter in subtidal sandy sediments (mediterranean sea) davide moccia,* alessandro cau, maria carmela meloni, antonio pusceddu department of life and environmental sciences, university of cagliari, via fiorelli 1, 09126 cagliari, italy abstract while variations in sedimentary organic matter (om) quantity, biochemical composition and nutritional quality as well as in meiofaunal abundance and assemblage composition at the macroand mesoscale are relatively well known, information about variations at the microscale is much scarcer. to shed some light on this issue, we tested the null hypothesis by which abundance and composition of the meiofaunal assemblages, and the quantity, biochemical composition and nutritional quality of sedimentary organic matter in coastal shallow environments do not vary within a frame of 1 m2. no significant variation within the frame emerged for om quantity, nutritional quality, biochemical composition and the abundance of meiofaunal assemblages. on the other hand, the composition of meiofaunal assemblages varied significantly within the frame and exhibited a clear segregation of assemblages farther to the shore, as a likely result of local micro-hydrodynamic conditions. spatial autocorrelation analysis revealed that lipid and protein sedimentary contents had a random distribution, whereas carbohydrate and biopolymeric c contents and meiofaunal total abundance were characterized by a patchy distribution, with discrete peaks within the sub-frame squares (ca. 0.1 m2). phytopigments showed a spatial positive autocorrelation distribution, following the micro-hydrodynamic pattern, with patches larger than the sub-frame square, but smaller than the entire one (1 m2). overall, our results suggest that, within 1 m2 of subtidal sandy sediments, three replicates could be sufficient to assess correctly om attributes and the abundance of meiofauna, but could be possibly inadequate for assessing meiofaunal assemblages’ composition at a finer scale (<1 m2). no nco mm er cia l u se on ly d. moccia et al.58 to provide further insights on the spatial scales of variation of meiofauna, we analyzed the small-scale (<1m2) variation in the abundance and composition of the meiofaunal assemblages, along with the om quantity, biochemical composition and nutritional quality in subtidal sandy sediments. more in details, we tested the null hypothesis by which abundance and composition of the meiofaunal assemblage, om quantity, biochemical composition and nutritional quality do not vary within a surface of 1 m2. finally, we investigated the small-scale spatial autocorrelation of the organic compounds in terms of phytopigment, protein, carbohydrate, lipid, biopolymeric c (bpc) contents and meiofaunal abundance to: i) analyze the heterogeneity and variability of the spatial distribution at their smallest spatial scale of distribution; ii) characterize their spatial patterns of variation, as an indication of the reliability of the sample size and replication. methods sediment sampling was carried out at 50-cm depth in a coastal sandy location of south sardinia (tyrrhenian sea, mediterranean sea) in may 2016. a 1 × 1 m frame was placed on the sea bottom and divided into 3 × 3 squares (each one of 33 × 33 cm size) for a total of 9 subframes (fig. 1). from each of the 9 sub-frames, 6 replicated sediment cores (randomly placed) were collected manually using plexiglass corers (4.7 cm internal diameter), of which three of them dedicated to the analysis of sediment organic matter, and three to the analysis of the meiofauna. for the om determinations, the first 2 cm of each sediment core were stored in petri dishes at -20°c until the analysis. protein, carbohydrate and lipid contents were determined spectrophotometrically according to danovaro (2010). for each biochemical assay, blanks were obtained using pre-calcinated sediments (450°c for 4 h). all the analyses were performed in triplicate, with about 1 g of sediment per replicate. carbohydrate, protein and lipid sedimentary contents were converted into carbon equivalents using the conversion factors of 0.40, 0.49 and 0.75 mg c mg–1, respectively, and their sum defined as biopolymeric c (bpc) (fabiano et al., 1995). phytopigments were extracted using 5 ml of 90% acetone (at 4°c in the dark for 12 h) from 0.5 g sediment samples (lorenzen and jeffrey, 1980). concentrations of chlorophyll-a and phaeopigments, after acidification of extracts with 200 ml 0.1 n hcl, were determined fluorometrically (danovaro, 2010). total phytopigments concentrations were calculated as sum of chlorophyll-a and phaeopigment concentrations. the algal c contribution to bpc was calculated as the percentage of phytopigment-to-bpc concentrations after converting the total phytopigment concentrations in to c equivalents using a mean value of 40 mg c mg–1. this, the protein contribution to bpc and the protein to carbohydrate ratio were used as proxies for om nutritional quality (pusceddu et al., 2009, 2010). the meiofauna was extracted by decantation (heip et al., 1985) as the sediments were dominated by sand. sediments were sieved through a 40 mm mesh. the filtered material was collected and stored in 50 ml jars, diluted with marine water and fixed by buffered formalin (ph 8.2, 2% final solution; higgins and thiel, 1988). the major meiobenthic organisms were counted and classified per taxon under a stereomicroscope (25-50x magnification) after staining with rose bengal (0.5 g l–1). the number of individuals obtained from each core was normalized to 10 cm2. a non-parametric permutational analyses of variance (permanova; anderson, 2001) was performed to test for differences in organic matter quantity, biochemical composition and nutritional quality, as well as meiofauna abundance and taxonomic composition, within the 1 m2 frame. two different designs were used: i) considering samples station randomly distributed within the frame as single source of variation; ii) considering two main sources of variation: transect as fixed factor (t =3 levels, parallel to the coast line and fig. 1. sampling design: the frame utilized in the present study is divided into 3 transects: a, b, and c, with transect a on the coast side and transect c towards the sea. numbers indicate the three stations in each of the three transect. no nco mm er cia l u se on ly perpendicular to the prevalent wave motion) and stations (s=9) as random factor nested in t. the analysis of the om dataset was based on euclidean distances resemblance matrix of previously normalized data. the analysis of the meiofaunal dataset was based on: i) bray-curtis similarity matrix of either untransformed or presence/absence transformed data. the visual representation of the differences between stations/transects was obtained by a non-metric multidimensional scaling (nmds). the use of presence/absence transformed data was chosen in order to scaling down the importance of highly abundant taxa (like nematodes) and, therefore, giving the same importance in variations to rare taxa (anderson et al., 2001). simper analysis was performed to assess the percentage dissimilarity in the meiofaunal taxonomic composition between transects and sampling station to identify which species contributed most to the observed dissimilarities. the relative influence of the different biochemical compounds on variations in the composition of meiofaunal assemblages was investigated through a nonparametric multivariate multiple regression analysis using the routine distlm, with either untransformed and presence/absence transformed data. all statistical analyses were performed through the use of the software primer 6+, using the routine included in the package permanova (anderson et al., 2008). small-scale dispersion analyses of sedimentary organic components (phytopigment, protein, carbohydrate, lipid and bpc concentrations) and meiofauna, were conducted using spatial autocorrelation techniques (cliff and ord, 1973; sokal and oden, 1978) to evaluate whether the observed spatial pattern of a variable is either random or aggregated so that the concentration/abundance value at one station is dependent on values present in neighboring station. in this analysis, moran’s index, geary’s coefficient (weighting of distance−2, jumars et al., 1977) and fisher’s index (s2/x) were used to analyze spatial patterns. moran’s i, ranging between −1 (maximum negative autocorrelation) and +1 (maximum positive autocorrelation) was used to detect aggregation due to extreme values in adjacent cores. geary’s coefficient, ranging from 0 (maximum positive autocorrelation) to a positive value (>1) for negative autocorrelation was used to test whether adjacent cores contain similar values (jumars et al., 1977). when i and c are both significantly >0 the distribution reflects plain positive spatial autocorrelation, whereas when only i is significant, the distribution creates discrete peaks in abundance. furthermore, if i, c, and the fisher’s index are all not significant, the distribution is random (pinckney and sandulli, 1990), whereas if fisher’s index is significant but i and c are not, it means that the size of the aggregates is less than the sampling scale (sokal and wartenberg, 1981). results organic matter spatial variability chlorophyll-a, phaeopigment, total phytopigment, protein, carbohydrate, lipid and biopolymeric c contents, as well as the algal and protein contributions to biopolymeric c and the protein to carbohydrate ratio, at all sampling stations are given in tab. 1. the permutational analysis of variance was first ran considering sampling stations as unique source of variation. none of the organic compounds showed significant spatial variations (tab. 2). the biopolymeric c content varied from 125.89±22.80 to 186.92±53.85 µg c g–1 (mean=154.95±28.11 µg c g–1; fig. 2a). carbohydrates (ranging from 217.87±36.05 to 271.55±25.15 µg g–1) were the dominant organic compound with an average 63% contribution to the biopolymeric c, followed by proteins (ranging from 68.33±16.95 to 130.57±69.63 µg g–1, 34% of bpc on average), and lipids (ranging from 3.37±0.81 to 10.33±0.71 mg g–1; 3% of bpc on average, fig. 2b). total phytopigments concentration (as the sum of chlorophyll-a and phaeopigments) varied from 0.61±0.06 to 1.00±0.49 µg g–1, with a general dominance of chlorophyll-a (up to 0.77±0.36 µg g–1, average of 79% of total phytopigments) over phaeopigments (maximum value 0.23±0.13, average of 21%) (fig. 2c ). biochemical composition and nutritional quality of sedimentary organic matter did not vary among stations as a whole (tab. 2). spatial variability of meiofauna overall, a total of 9 major taxa were found within the 1 m–2 sampling frame: nematoda, amphipoda, copepoda, gastropoda, gastrotricha, oligochaeta, ostracoda, polychaeta and tardigrada, six out of which were ubiquitous in all the sampling stations (nematodes, amphipods, copepods, gastropods, gastrotriches and ostracods). polychaetes were only encountered within stations along the transects a and b, oligochaetes were encountered exclusively in stations along transect b and tardigrades were exclusively encountered in stations along transect a (tab. 3). the total abundance of meiofauna showed no significant variation within the frame (tab. 4a; fig. 3a). differently, the composition of the meiofaunal assemblages (using either untransformed or presence/absence transformed data) significantly varied within the frame when sample stations were considered as the only source of variation (tab. 4a; fig. 3b). when two factors of variation were considered, transects (a, b, c) and stations (1,2,3), the analysis of variation based on untransformed data showed the presence of significant spatial variations only among transects (tab. 4b). when the analysis was conducted on presence/absence transformed data, the differences were significant only among stations within each transect (tab. 4b). a higher nematode dominance (avsmall-scale distribution of coastal meiofaunal assemblages and sedimentary organic matter 59 no nco mm er cia l u se on ly d. moccia et al.60 erage of 90% of the total meiofaunal abundance) occurred in transect a, when compared to the other two transects (ca. 85% in both transect b and c). the abundance of copepods and gastrotrichs increased, from an average of 5% and 0.1% in transect a, to an average of 10% and 0.8% in transect c, respectively (fig. 3b). this pattern was further confirmed by the simper analysis, that reveals that differences in the relative abundance of nematodes and copepods were the most responsible for the observed spatial dissimilarities. the simper analysis also reveals that the larger dissimilarities in the composition of meiofaunal assemblages occurred between transect c and the other two transects (tab. 5). variation among the three transects was also confirmed by the nmds bi-plot, that shows a clear visual segregation between transect c and the other two transects (fig. 4). ta b. 1 .s ed im en ta ry o rg an ic m at te r c on te nt s an d nu tr iti on al q ua lit y in s an dy s ed im en ts o f t he s tu dy s ite . st at io n c hl or op hy lla p ha eo pi gm en t to ta l p ig m en ts p ro te in c ar bo hy dr at e l ip id b p c p ro te in : b p c a lg al : b p c p ro te in : c ar bo hy dr at e (µ g g– 1 ) (µ g g– 1 ) (µ g g– 1 ) (µ g g– 1 ) (µ g g– 1 ) (µ g g– 1 ) (µ g c g –1 ) ( % ) ( % ) a 1 0. 55 ±0 .0 1 0 .1 5± 0. 02 0. 70 ±0 .0 1 1 30 .5 7± 69 .6 3 2 71 .5 5± 25 .1 5 7. 77 ±0 .3 6 1 86 .9 2± 53 .8 5 3 7. 00 ±1 2. 59 1 1. 82 ±3 .5 3 0. 58 ±0 .2 2 a 2 0. 44 ±0 .1 2 0 .2 2± 0. 14 0. 66 ±0 .0 3 98 .3 8± 16 .0 4 28 9. 21 ±9 1. 58 3. 37 ±0 .8 1 1 66 .4 2± 36 .2 0 29 .7 5± 6. 79 12 .4 2± 3. 13 0. 36 ±0 .1 1 a 3 0. 56 ±0 .0 6 0 .1 5± 0. 03 0. 71 ±0 .1 0 1 11 .6 4± 18 .8 5 2 20 .6 5± 38 .6 5 4. 45 ±0 .8 2 1 46 .3 0± 21 .8 1 37 .4 0± 3. 38 14 .6 8± 2. 33 0. 50 ±0 .0 8 b 1 0. 62 ±0 .0 7 0 .1 8± 0. 01 0. 80 ±0 .0 7 68 .3 3± 16 .9 5 21 7. 87 ±3 6. 05 7. 01 ±2 .1 0 1 25 .8 9± 21 .8 0 26 .4 7± 2. 65 19 .6 6± 4. 35 0. 31 ±0 .0 4 b 2 0. 77 ±0 .3 6 0 .2 3± 0. 13 1. 00 ±0 .4 9 91 .1 1± 17 .9 7 22 3. 52 ±1 8. 84 7. 56 ±1 .5 1 1 39 .7 2± 11 .4 8 31 .8 5± 4. 57 2 1. 57 ±1 0. 36 0 .4 1± 0. 09 b 3 0. 73 ±0 .3 3 0 .1 1± 0. 10 0. 84 ±0 .2 5 1 25 .1 8± 19 .2 1 2 27 .6 4± 44 .8 1 4. 60 ±0 .8 8 1 55 .8 4± 15 .5 2 39 .6 0± 7. 01 16 .2 1± 4. 65 0. 57 ±0 .1 7 c 1 0. 70 ±0 .1 4 0 .1 5± 0. 03 0. 85 ±0 .1 6 1 23 .2 1± 59 .9 0 2 44 .5 3± 46 .6 2 5. 94 ±2 .2 2 1 62 .6 4± 15 .4 7 3 6. 38 ±1 5. 14 1 5. 67 ±2 .3 8 0. 55 ±0 .3 8 c 2 0. 47 ±0 .0 6 0 .1 4± 0. 04 0. 62 ±0 .0 3 92 .3 8± 18 .0 1 22 8. 50 ±7 0. 87 8. 60 ±0 .9 2 1 43 .1 1± 35 .4 6 32 .2 4± 5. 24 13 .6 8± 4. 48 0. 42 ±0 .1 1 c 3 0. 55 ±0 .0 3 0 .0 6± 0. 05 0. 61 ±0 .0 6 1 25 .4 9± 73 .8 8 2 46 .0 9± 29 .4 9 1 0. 33 ±0 .7 1 1 67 .6 7± 40 .3 9 3 4. 94 ±1 2. 11 1 1. 28 ±2 .2 8 0. 50 ±0 .2 9 b pc , b io po ly m er ic c . fig. 2. sedimentary organic matter contents and biochemical composition in the investigated sediments. a) biopolymeric c content, b) protein, carbohydrate and lipid contribution to biopolymeric c, c) chlorophyll-a, phaeopigment and total phytopigment contents. no nco mm er cia l u se on ly small-scale distribution of coastal meiofaunal assemblages and sedimentary organic matter 61 organic matter-meiofauna relationships the results of the distlm forward analysis reveals that lipids and phaeopigments explained 37% and 22%, respectively, of meiofaunal assemblage variations, when the analysis was run on untransformed data, whereas, when the analysis was carried out on presence/absence data, lipids and proteins explained 41% and 27% of variation. tab. 2. results of the permanova testing for differences among stations in the composition and the nutritional quality of sedimentary organic matter in 1 m2 of subtidal sandy sediments. reported is also the percentage of variance explained by the source of variation (station). variable source df ms pseudo f p (mc) % of explained variance protein station 8 0.884 0.841 ns 60 carbohydrate station 8 1.117 1.179 ns 6 lipid station 8 1.806 2.813 ns 38 biopolymeric c station 8 1.309 1.517 ns 15 chlorophyll-a station 8 1.194 1.307 ns 9 phaeopigment station 8 1.231 1.372 ns 11 total phytopigments station 8 1.212 1.339 ns 10 protein: carbohydrate station 8 0.625 0.533 ns 0 protein fraction of bpc station 8 0.742 0.664 ns 0 algal fraction of bpc station 8 1.481 1.882 ns 0 biochemical composition station 8 5.016 1.416 ns 0 df, degrees of freedom; ms, means square; f, statistic f; p (mc), monte carlo probability level; bpc, biopolymeric c; ns, not significant. tab. 3. meiofaunal abundance (numbers of individuals per 10 cm2 ± standard error) in the sandy sediment of the study site. station nematodes amphipods copepods gasteropods gastrotrichs oligochaetes ostracods polychaetes tardigrades a1 2247±187 91±38 122±31 5±3 5±2 0 5±33 3±2 0 a2 2626±446 106±13 128±18 17±6 3±2 0 14±5 1±1 1±0 a3 2302±133 167±15 168±26 13±5 1±1 0 13±6 0 0 b1 2920±121 170±21 154±4 21±3 2±2 2±0 4±1 0 0 b2 1875±197 138±31 103±23 13±3 12±1 0 3±1 2±2 0 b3 2146±336 140±32 139±42 8±2 22±7 1±1 3±1 1±1 0 c1 2611±231 161±28 310±29 3±3 36±4 0 4±2 0 0 c2 2109±204 121±13 253±10 1±1 20±5 0 3±2 0 0 c3 3612±386 128±16 238±15 0±0 15±1 0 3±1 0 0 tab. 4. results of the permanova testing for: a) differences among stations (one factor design) and b) differences among transects (tr) and stations (st) (two factors design) in the abundance and composition (using untransformed and presence/absence transformed data, respectively) of the meiofaunal assemblages. also reported is the percentage of variance explained by each source of variation. a) one factor design variable source df ms pseudo f p(mc) % of explained variance total meiofaunal abundance station 8 165.4 2.037 ns 24 assemblage composition(untransformed data) station 8 202.5 2.096 * 29 community composition (presence/absence transformed data) station 8 214.3 3.165 ** 41 b) two factors design variable source df ms pseudo f p(mc) % of explained variance assemblage composition (untransformed data) tr 2 342.9 3.596 * 31 st (tr) 6 95.4 2.263 ns 20 assemblage composition (presence/absence transformed data) tr 2 370.8 1.909 ns 28 st (tr) 6 194.3 3.005 ** 34 df, degrees of freedom; ms, means square; f, statistic f; p (mc), monte carlo probability level; bpc, biopolymeric c; ***p<0.001; **p <0.01; *p<0.05, ns, not significant. no nco mm er cia l u se on ly d. moccia et al.62 spatial autocorrelation analyses and surface plots the results of the autocorrelation analysis are reported in tab. 6, moran’s i, geary’s c, and fisher’s index were determined for all variables. results showed that within 1 m2 sampling area, proteins and lipids showed a random spatial distribution (i, c and fisher’s index p>0.05, tab. 6), whereas carbohydrates, biopolymeric c and meiofaunal total abundance exhibited a patchy distribution, with patches smaller than 1 m2, since i and c do not depart significantly from 1, but fisher’s index is significantly >1 (tab. 6). only chlorophyll-a and phaeopigment showed a positive auto-correlated pattern (i=0.34, p < 0.05 c=0.52, p < 0.05; i= 0.34 p< 0.05 c=0.58 p<0.05) with patch size larger than the single sampling square (0.1 m2), but lower than the total sampling area (1 m2) (tab. 6). discussion and conclusions a detailed understanding of the spatial distribution patterns of consumers and their resources and the scale at which their interactions occur is essential to assess a correct and reliable size of samples (sandulli and pickney, 1999, semprucci et al., 2011). an inappropriate sample size, indeed, may lead to underor over-estimate spatial heterogeneity and, thus, to formulate erroneous conclusions (cerca et al., 2018). in this study, we analyzed the small-scale variation of the meiofauna assemblages (in terms of either abundance and taxonomic composition) and the quantity, nutritional quality and biochemical composition of om in subtidal sediments, to test the hypothesis by which meiofaunal assemblages and om quantity and composition do not vary within a surface of 1 m2. our results show that sediments under scrutiny were characterized by a generally homogeneous spatial distrifig. 3. total meiofauna abundance (a) and assemblage composition (b) in transects and stations of the investigated sediments. fig. 4. nmds ordination of meiofauna assemblage data in the investigated sediments. no nco mm er cia l u se on ly small-scale distribution of coastal meiofaunal assemblages and sedimentary organic matter 63 bution of the trophic conditions, in terms of om quantity, biochemical composition and nutritional quality. a similar homogeneous distribution pattern at the small scale (i.e., within 1 m2) emerged also for the meiofaunal total abundance. these results, corroborated by the lack of autocorrelation patterns, indicate that the collection of 3 random replicas within a 1 m2 frame can be enough to represent reliably the mean characteristics of either om or meiofaunal abundance within such a spatial scale. on the other hand, the significant variation (permanova p<0.05) of the meiofaunal community composition among the three transects at this small spatial scale (1 m2), indicates that the sampling effort for assessing the variability of meiofaunal assemblage composition in sandy shallow sediments exposed to wave action would need a higher number of (possibly larger) replica per single m2. in this regard, we showed that changes in the composition of the meiofaunal assemblages across transects were associated with variations in the micro-hydrodynamic gradient, which decreased from transect a (nearest to the coast) and the transect c. although. our experiment is limited to the spring season, and thus not exportable on tab. 5. dissimilarities in the composition of meiofaunal assemblages among transects and stations and species responsible for the estimated difference. dissimilarity (%) explanatory species explained variance (%) cumulative explained variance (%) transect a vs b 10.50 nematodes, amphipods 34.79, 14.70 49.49 transect a vs c 12.64 nematodes, copepods 26.30, 22.56 48.48 transect b vs c 12.49 nematodes, copepods 33.14, 24.89 58.03 a1 vs a2 16.06 polychaetes, gasteropods 28.55, 18.35 46.90 a1 vs a3 18.50 gastrotrichs, polychaetes 32.95, 32.11 65.05 a2 vs a3 10.45 gastrotrichs, tardigrades 45.91, 28.19 74.10 a1 vs b1 25.08 oligochaetes, gastrotrichs 32.52, 22.26 54.78 b2 vs b1 17.54 oligochaetes, gastrotrichs 45.13, 25.17 70.30 a3 vs b1 12.30 oligochaetes, gastrotrichs 68.52, 31.48 100 a1 vs b2 10.48 polychaetes, gasteropods 50.00, 34.19 84.19 a2 vs b2 8.83 polychaetes, gastrotrichs 38.71, 30.65 69.35 a3 vs b2 8.60 gastrotrichs, polychaetes 68.52, 31.48 100 b1 vs b2 15.83 oligochaetes, gastrotrichs 50.00, 34.14 84.19 a1 vs b3 12.66 polychaetes, gasteropods 35.72, 22.70 58.43 a2 vs b3 10.92 polychaetes, gastrotrichs 30.20, 24.24 54.44 a3 vs b3 10.76 oligochaetes, gastrotrichs 53.46, 23.27 76.73 b1 vs b3 13.03 oligochaetes, gastrotrichs 41.54, 40.60 82.14 b2 vs b3 5.62 polychaetes, oligochaetes 58.61, 41.39 100 a1 vs c1 12.85 polychaetes, gasteropods 44.86, 30.78 75.64 a2 vs c1 11.31 gasteropods, gastrotrichs 26.06, 25.32 51.37 a3 vs c1 9.49 gastrotrichs, gasteropods 65.69, 34.04 100 b1 vs c1 17.03 oligochaetes, gastrotrichs 49.09, 33.61 82.70 b2 vs c1 5.58 gasteropods, polychaetes 52.78, 47.22 100 b3 vs c1 7.76 gasteropods, oligochaetes 37.05, 31.47 68.53 a1 vs c2 14.05 polychaetes, gasteropods 42.26, 28.87 71.13 a2 vs c2 14.49 gasteropods, gastrotrichs 20.33, 20.33 40.66 a3 vs c2 12.93 gastrotrichs, gasteropods 50.00, 25.00 75.00 b1 vs c2 20.38 oligochaetes, gastrotrichs 42.19, 28.91 71.09 b2 vs c2 8.60 gasteropods, ostracods 34.26, 34.26 68.52 b3 vs c2 10.76 gasteropods, ostracods 26.73, 26.73 53.46 c1 vs c2 7.27 gasteropods, ostracods 56.94, 43.06 100 a1 vs c3 15.25 polychaetes, gasteropods 40.07, 38.08 78.15 a2 vs c3 17.68 gasteropods, gastrotrichs 50.00, 17.14 67.14 a3 vs c3 16.36 gasteropods, gastrotrichs 59.26, 40, 74 100 b1 vs c3 23.74 gasteropods, oligochaetes 37.23, 37.23 74.47 b2 vs c3 11.62 gasteropods, polychaetes 76.09, 23.91 100 b3 vs c3 13.75 gasteropods, oligochaetes 62.71, 18.64 81.36 c1 vs c3 6.06 gasteropods 100 100 c2 vs c3 9.70 gasteropods, ostracods 65.63, 34.38 100 no nco mm er cia l u se on ly d. moccia et al.64 wider temporal scales, we notice that, across such gradient, the abundance of copepods and amphipods increased, whereas that of nematodes decreased. such a pattern is consistent with the habitus of these taxa (higgins and thiel, 1988; thiel, 1992; giere, 2009): copepods have a preeminent epi-benthic habitus, so that they are more sensitive to higher hydrodynamic conditions (where waves break), where, instead, nematodes, with a preeminent infaunal habitus, can easily dig the sediment to resist to waves’ disturbance. moreover, several nematodes species are capable of adhering to the sediment particles thanks to a gland that secretes sticky mucus, forming an “armature” that increases their body weight and, thereby, their resistance to the hydrodynamic disturbance (moens et al., 2013). the response of nematodes to hydrodynamic action (waves and currents) have been observed also in tropical areas (maldives), in which the effect of the higher exposition to the monsoon shaped the meiofaunal assemblages showing a higher nematodes abundance in exposed areas compared to those exposed to from low to medium physical disturbance (semprucci et al., 2010, 2011). although the sediment grain size variation within 1 m2 in a microtidal area, like the one under scrutiny can be assumed to be invariant, another potential bias of our results remains the lack of data about sediment grain size, which typically responds to physical disturbance and, in turn, represents a major factor controlling distribution of meiofauna (semprucci et al., 2010). the analyses carried out to identify the type of spatial pattern (random or patchy) of the investigated variables indicate that carbohydrate concentration, bpc and total meiofaunal abundance showed a patchy distribution, as previously reported also for bacteria and other biopolymers in sandy sediments of the adriatic sea (danovaro et al., 2001). the significant fisher’s index indicates also that for all of those variables the patch size was smaller than the surface of two adjacent squares (ca. <0.2 m2). on the other hand, chlorophyll-a and phaeopigments (which typically represent the most important nutritional component of sedimentary om, even in the deep sea; bianchelli et al., 2008; pusceddu et al., 2009) revealed a patchy distribution characterized by patch sizes higher than a single sub-frame square (i.e., >0.1 m2). in particular, chlorophyll-a sedimentary contents slightly increased while moving far from the shore. one possible explanation for such a pattern is that microalgal patch size might be influenced by the micro-scale hydrodynamic features which, in turn, shape the bottom micro-topography. this latter hypothesis is consistent with previous findings from micro-scale distribution analyses (danovaro et al., 2001), that identified sandy ripples and marks as the sedimentary structures shaped by waves motion in shallow waters that are most responsible for the micro-distribution patterns of benthic phytopigments. we stress that the microalgal patch size reported in this study (<0.1 m2) could be up to two orders of magnitude larger than that reported in adriatic sediments (1.34 cm2). we notice, however, that such a huge discrepancy could be due to the much different size of the sub-frame square used in this study when compared to that used by danovaro et al., (2001), who analyzed samples collected centimeters one a part to assess the micro-topographical significance of ripplesmarks structures and the depression among them. many studies have shown that the composition of meiofaunal communities depends upon the quantity and biochemical composition of sedimentary organic matter over a broad range of spatial scales and habitats (cerrano et al., 2010; pusceddu et al., 2007, 2011; semprucci et al., 2010; bianchelli et al., 2013). among the different classes of organic compounds, phytopigments, proteins and lipids often represent high-energy components of the biopolymeric c (pusceddu et al., 2009) and, indeed, have been often related to meiofaunal abundance and biodiversity either in shallow (albertelli et al., 1999) and deepsea (danovaro et al., 1999; gambi et al., 2014) habitats. most typically, algal and protein fractions of biopolymeric c represent the most bioavailable components for benthic consumers nutrition (pusceddu et al., 2003). accordingly, tab. 6. results of spatial autocorrelation analyses of all variables. moran’s i, geary’s c, and fisher’s index are reported. autocorrelation values were calculated using weighting of distance−2 and significant values were assigned using the randomization assumption. expected values e (i)=−0.020 and e (c)=1.000. variable moran’s i geary’s c fisher’s index proteins -0.33 ns 1.25 ns 4.12 ns carbohydrates -0.06 ns 0.80 ns 102.66 *** lipid -0.16 ns 0.96 ns 0.75 ns biopolymeric c -0.13 ns 1.00 ns 7.09 *** chlorophyll-a 0.34 * 0.52 * 0.02 ns phaeopigments 0.34 * 0.58 * 0.01 ns meiofauna -0.26 ns 1.08 ns 62.3 *** *p<0.05; **p<0.01; ***p<0.001; ns, not significant. no nco mm er cia l u se on ly small-scale distribution of coastal meiofaunal assemblages and sedimentary organic matter 65 the results of the distlm analysis carried out in this study suggest that the spatial variation of meiofauna even at the small scale considered here is tightly related to the concentration of lipids and phaeopigments (meiofaunal abundance) or proteins (community composition). corresponding author: mocciadavide@unica.it key words: biochemical composition; nutritional quality; spatial variability; sampling size; coastal content; trophic status. received: 14 march 2019. accepted: 21 june 2019. this work is licensed under a creative commons attribution noncommercial 4.0 license (cc by-nc 4.0). ©copyright: the author(s), 2019 licensee pagepress, italy advances in oceanography and limnology, 2019; 10:8169 doi: 10.4081/aiol.2019.8169 references albertelli g, covazzi-harriague a, danovaro r, fabiano m, fraschetti s, pusceddu a, 1999. differential responses of bacteria, meiofauna and macrofauna in a shelf area (ligurian sea, nw mediterranean): role of food availability. j. sea. res. 42:11-26. anderson mj, 2001. permutation tests for univariate or multivariate analysis of variance and regression. can. j. fish. aquat. sci. 58:626–639. anderson mj, gorley rn, clarke kr, 2008. permanova+ for primer: guide to software and statistical methods. primer-e: plymouth, uk. bianchelli s, gambi c, pusceddu a, danovaro r, 2008. trophic conditions and meiofaunal assemblages in the bari canyon and the adjacent open slope (adriatic sea). chem. ecol. 24:101-109. bianchelli s, pusceddu a, canese s, greco s, danovaro r, 2013. high meiofaunal and nematodes diversity around mesophotic coral oases in the mediterranean sea. plos one 8:e66553. blanchard gf, 1990. overlapping microscale dispersion patterns of meiofauna and microphytobethos. mar. ecol. prog. ser. 68:101-111. cerca j, purschke g, struck th, 2018. marine connectivity dynamics: clarifying cosmopolitan distributions of marine interstitial invertebrates and the meiofauna paradox. mar. biol. 165:1-21. cerrano c, danovaro r, gambi c, pusceddu a, riva a, schiaparelli s, 2010. gold coral (savalia savaglia) and gorgonian forests enhance benthic biodiversity and ecosystem functioning in the mesophotic zone. biodivers. conserv. 19:153-167. chandler gt, fleeger jw, 1987. facilitative and inhibitory interactions among estuarine meiobenthic harpacticoid copepods. ecology 68:1906-1919. cliff a, ord j, 1973. spatial autocorrelation. london: pion. danovaro r, 2010. methods for the study of deep-sea sediments, their functioning and biodiversity. crc press, taylor & francis group, boca raton. danovaro r, armeni m, dell’anno a, fabiano m, manini e, marrale d, pusceddu a, vanucci s, 2001. small-scale distribution of bacteria, enzymatic activities and organic matter in coastal sediments. microb. ecol. 42:177-185. danovaro r, dinet a, duineveld g, tselepides a, 1999. benthic response to particulate fluxes in different trophic environments: a comparison between the gulf of lions – catalan sea (western-mediterranean) and the cretan sea (eastern-mediterranean). prog. oceanogr. 44:287-312. decho aw, hummon wd, fleeger jw, 1985. meiofaunasediment interactions around subtropical seagrass sediments using factor analysis. j. mar. res. 43:237-255. dell’anno a, mei ml, pusceddu a, danovaro r, 2002. assessing the trophic state and eutrophication of coastal marine systems: a new approach based on the biochemical composition of sediment organic matter. mar. pollut. bull. 44:611-622. denny mw, helmuth b, leonard gh, harley cdg, hunt ljh, nelson ek, 2004. quantifying scale in ecology: lessons from a wave-swept shore. ecol. monogr. 74;513-532. fabiano m, danovaro r, fraschetti s, 1995. a three-year time series of elemental and biochemical composition of organic matter in subtidal sandy sediments of the ligurian sea (northwestern mediterranean). cont. shelf. res. 15:14531469. fleeger jw, decho aw, 1987. spatial variability of interstitial meiofauna: a review. stygologia 3:35-54. fontaneto d, hortal j, 2012. microbial biogeography: is everything small everywhere?, p. 87-98. in: in: a. lesley, r. ogilvie and p. hirsch (eds.), microbial ecological theory current perspectives. caister academic press. fontaneto d, 2019. long-distance passive dispersal in microscopic aquatic animals. movem. ecol. 7:10. gambi c, pusceddu a, benedetti-cecchi l, danovaro r, 2014. species richness, species turnover and functional diversity in nematodes of the deep mediterranean sea: searching for drivers at different spatial scales. global. ecol. biogeogr. 23:24-39. heip c, vincx m, vranken g, 1985. the ecology of marine nematodes. oceanogr. mar. biol. 23:399-489. hewitt je, mcbride gb, pridmore rd, thrush sf, 1993. patchy distributions: optimizing sample size. env. monit. assess. 27:95-105. higgins rp, thiel h, 1988. introduction to the study of meiofauna. smithsonian institution press, washington, dc. jumars m, thistle d, jones m, 1977. detecting twodimensional spatial structure in biological data. oecologia 28:109-123. legendre p, fortin mj, 1989. spatial pattern and ecological analysis. vegetatio 80:107-138. levin la, sibuet m, 2012. understanding continental margin biodiversity: a new imperative. annu. rev. mar. sci. 4:79112. lorenzen c, jeffrey j, 1980. determination of chlorophyll in seawater. unesco technical paper in marine science no. 35. unesco, paris. moens t, braeckman u, derycke s, fonseca g, gallucci f, no nco mm er cia l u se on ly d. moccia et al.66 ingels j, leduc d, vanaverbeke j, van colen c, vanreusel a, vinx m, 2013. ecology of free-living marine nematodes, p. 109-152 in: a. schmidt-rhaesa (ed.), handbook of zoology: gastrotricha, cycloneuralia and gnathifera, vol. 2: nematoda. de gruyter, berlin. montagna pa, spies rb, 1985. meiofauna and chlorophyll-a associated with beggiatoa mats of a natural submarine petroleum seep. mar. environ. res. 16:231-242. pinckney j, sandulli r, 1990. spatial autocorrelation analysis of meiofaunal and microalgal populations on an interdital sandflat: scale linkage between consumers and resources. estuar. coastal shelf. sci. 30:341-353. pratt d, pilditch c, lohrer a, thrush s, kraan c, 2015. spatial distributions of grazing activity and microphytobenthos reveal scale-dependent relationships across a sedimentary gradient. estuar. coast. 38:722-734. pusceddu a, bianchelli s, canals m, sanchez-vidal a, durrieu de madron x, heussner s, lykousis v, de stigter h, trincardi f, danovaro r, 2010. organic matter in sediments of canyons and open slopes of the portuguese, catalan, southern adriatic and cretan sea margins. deep-sea res. pt i 57:441-457. pusceddu a, bianchelli s, gambi c, danovaro r, 2011. assessment of benthic trophic status of marine coastal ecosystems: significance of meiofaunal rare taxa. estuarine coastal shelf. sci. 93:420-430. pusceddu a, dell’anno a, fabiano m, danovaro r, 2004. quantity and biochemical composition of organic matter in marine sediments. biol. mar. medit. 11(s1):39-53. pusceddu a, dell’anno a, fabiano m, danovaro r, 2009. quantity and bioavailability of sediment organic matter as signatures of benthic trophic status. mar. ecol. progr. ser. 375:41-52. pusceddu a, dell’anno a, manini e, fabiano m, sarà g, danovaro r, 2003. enzymatically hydrolyzable protein and carbohydrate sedimentary pools as indicators of the trophic state of ‘detritus sink’ systems: a case study in a mediterranean coastal lagoon. estuaries 26:641-650. pusceddu a, gambi c, corinaldesi c, scopa m, danovaro r. 2014. relationships between meiofaunal biodiversity and prokaryotic heterotrophic production in different tropical habitats and oceanic regions. plos one 9(3):e91056. pusceddu a, gambi c, manini e, danovaro r, 2007. trophic state, ecosystem efficiency and biodiversity of transitional aquatic ecosystems: analysis of environmental quality based on different benthic indicators. chem. ecol. 23:1-11. raes m, de troch m, ndaro sgm, muthumbi a, guilini k, vanreusel a, 2007. the structuring role of microhabitat type in coral degradation zones: a case study with marine nematodes from kenya and zanzibar. coral reefs, 26: 113-126. rex m, etter kr, 2011. deep-sea biodiversity: pattern and scale. bioscience 61:327-328. sandulli r, pinckney j, 1999. patch sizes and spatial patterns of meiobenthic copepods and benthic microalgae in sandy sediments: a microscale approach. j. sea res. 41:179-187. semprucci f, boi p, manti a, covazzi harriague a, rocchi m, colantoni p, papa s, balsamo m, 2010. benthic communities along a littoral of the central adriatic sea (italy). helgol. mar. res. 64:101-115. sokal r, oden n, 1978. spatial autocorrelation in biology. ii. some biological implications and four applications of evolutionary and ecological interest. biol. j. linn. soc. 10:229-249. sokal rr, wartenberg de, 1981. space and population structure, p. 186-213. in: d. griffith and r. mckinnon (eds.), dynamic spatial models. sijthoff and noordhoff, aphen aan den rijn. thrush s, 1991. spatial patterns in soft-bottom communities. trends ecol. evol. 6:75–79. zeppilli d, pusceddu a, trincardi f, danovaro r, 2016. seafloor heterogeneity influences the biodiversity-ecosystem functioning relationships in the deep sea. sci. rep. 6:26352. warwick rm, gee jm, berge ja, ambrose w, 1986. effects of the feeding activity of the polychaete streblosoma bairdi (malmgren) on meiofaunal abundance and community structure. sarsia 71:11-16. woolley s, tittensor pd, dunstan kp, guillera-arroita g, lahoz-monfort j, wintle ba, worm b, o’hara t, 2016. deep-sea diversity patterns are shaped by energy availability. nature 533:393-396.no nco mm er cia l u se on ly layout 1 introduction despite intensive research identifying causes, consequences and possible preventive measures of cyanobacterial mass proliferations, cyanobacterial blooms represent a major problem in fresh waters throughout the world. the adverse consequences include water quality degradation, accumulation and microbial decay of bloom biomass followed by lowering of oxygen content in water (wiegand and pflugmacher, 2005). in addition, a wide spectrum of toxins and secondary metabolites produced by cyanobacteria have been shown to adversely affect aquatic organisms (codd et al., 2005; zanchett and oliveira-filho, 2013), livestock (mcgorum et al., 2015) and human health (kuiper-goodman et al., 1999; zanchett and oliveira-filho, 2013). while different cyanobacterial taxons can contribute to the formation of dense cyanobacterial water blooms depending on geographical and ecological conditions, microcystis sp. represents cosmopolitan and pervasive cyanobacterial genera, which is frequently reported to dominate water blooms in freshwaters of all continents except antarctica (harke et al., 2016). microcystis species (such as microcystis aeruginosa) are also among the prominent producers of the most broadly studied cyanobacterial toxins microcystins (mcs) (bláha et al., 2009). mcs are known liver tumour promoters (nishiwaki-matsushima et al., 1992), with the most common structural variant microcystin-lr (mc-lr) being classified by the international agency for research on cancer as a possible human carcinogen (iarc2b) (grosse et al., 2006). mcs have been shown to act via inhibition of serine/threonine protein phosadvances in oceanography and limnology, 2017; 8(1): 107-120 article doi: 10.4081/aiol.2017.6342 this work is licensed under a creative commons attribution-noncommercial 4.0 international license (cc by-nc 4.0). chlorination and ozonation differentially reduced the microcystin content and tumour promoting activity of a complex cyanobacterial extract iva sovadinová,1,2* pavel babica,1,2,3 ondřej adamovský,1,2 alla alpatova,4,5 volodymyr tarabara,4 brad luther upham,2 luděk bláha1 1recetox research centre for toxic compounds in the environment, faculty of science, masaryk university, kamenice 753/5, 62500 brno, czech republic; 2department of pediatrics and human development, and institute for integrative toxicology, michigan state university, 1129 farm lane, east lansing, mi 48824, usa; 3department of experimental phycology and ecotoxicology, institute of botany, czech academy of sciences, lidicka 25/27, 60200 brno, czech republic; 4department of civil and environmental engineering, michigan state university, 428 s. shaw lane, east lansing, mi 48824, usa; 5civil and environmental engineering, university of alberta, ab t6g 2e1 edmonton, canada *corresponding author: sovadinova@recetox.muni.cz abstract despite intensive research and management efforts in the past decades, cyanobacterial blooms and their toxins, such as microcystins (mcs), continue to represent a major ecological and health problem in fresh waters throughout the world. our objective was to compare the efficacy of two commonly used drinking water treatment technologies, chlorination and ozonation, in removing mcs and in reducing tumour promotion-related effects of cyanobacteria, such as inhibition of gap junctional intercellular communication (gjic) and activation of mitogen activated protein kinases (mapks) in a rat liver epithelial stem-like cell line (wb-f344). this combined chemical and bioassay approach demonstrated that ozone effectively removed all mcs from an extract of a globally important bloom-forming cyanobacterium, microcystis sp. ozone also significantly reduced the overall tumour promotional potency of the cyanobacterial extract, as indicated by a substantial reduction in the ability of the extract to inhibit gjic and activate extracellular receptor kinase 1/2 (erk1/2). although comparable reduction of total organic carbon was achieved by ozone and chlorine treatment, chlorination was much less effective in removing mcs and reducing the effects on gjic. chlorination had a biphasic effect with an observed decrease of extract-induced activation of erk1/2 at the lower chlorine doses; whereas at high doses of chlorine the by-products of chlorination actually induced the activation of erk1/2. the extracts induced p38 activation, and chlorination was not effective in reversing this effect, while ozone did reverse this effect, albeit not as much as the activation of erk1/2. thus, ozone was effective in reducing the toxicity of cyanobacterial extracts while chlorination was not only lacking efficacy, but at high doses of chlorination further produced by-products that were equally toxic as the untreated samples. our study indicates the value of using an effect-based approach to assess the efficacy of water treatment systems in removing toxins, and more specifically demonstrates that ozone was more effective at reducing the toxic potential of cyanobacterial-contaminated water. key words: cyanobacteria; water treatment; toxicity; chlorination; ozonation; microcystin. received: 11 october 2016. accepted: 14 february 2017.no nco mm er cia l u se on ly i. sovadinová et al.108 phatases (pps) (campos and vasconcelos, 2010) and induction of oxidative stress followed by damage to cellular macromolecules (mathe et al., 2016). chronic exposures to cyanobacteria and their toxins – e.g., via contaminated drinking water – have been associated with increased occurrence of liver and colorectal cancer (yu, 1995; zhou et al., 2002; svircev et al., 2009). other (often yet unidentified) compounds produced by complex cyanobacterial blooms can also induce or contribute to different adverse effects (oberemm et al., 1997; berry et al., 2009) including tumour promotion (bláha et al., 2010). occurrence of toxic cyanobacteria in surface water presents a challenge to drinking water treatment facilities. preventive measures as well as various water treatment technologies used to minimize human health risks caused by cyanobacteria and their toxins have been recently summarized and critically discussed (westrick et al., 2010; merel et al., 2013; roegner et al., 2014; hiskia et al., 2016; ibelings et al., 2016). the available approaches able to remove mcs to different extents include coagulationflocculation-sedimentation, standard oxidation and disinfection by chlorine or permanganate, ozonation and uv disinfection, sorption by activated carbon, nanoand ultrafiltration, and advanced oxidation processes (aop) (westrick et al., 2010; merel et al., 2013; roegner et al., 2014). different drinking water treatment technologies are applied in different countries and contexts, and evaluation of their treatment efficiency should focus on i) removal of targeted priority pollutants (e.g., mcs in case of cyanobacterial toxins) to comply with current treatment goals; as well as ii) removal of other potentially harmful and toxic components of the complex material which may not be fully chemically characterized; and iii) formation of new harmful metabolites/toxic by-products during the application of the water treatment technology (upham et al., 1994; upham et al., 1995; prasse et al., 2015). to address these different aspects of drinking water treatment, complementary chemical and biological tools (i.e. instrumental analyses and bioassays) should be included in the monitoring plans (maier et al., 2015). this ongoing effort can be highlighted by current implementation of the effect-based tools into the monitoring guidelines for water quality assessment (wernersson et al., 2014), which combines several bioassays targeting different toxic modes of action (moas) to provide additional information to classical chemical analyses and thus a more integrative view. gap junctional intercellular communication (gjic) plays a fundamental role in maintaining tissue homeostasis, and provides an excellent biological endpoint to assess potential adverse health effects of many anthropogenic toxicants and natural toxins (vinken et al., 2009). gjic is a critical cellular process for the coordination of different intra-, extra-, and inter-cellular signalling pathways required for proper cell behaviour, tissue development, tissue function and maintenance of tissue homeostasis. most chemical carcinogens and tumour promoters inhibit gjic in in vitro assays, and demonstrated to be a representative marker of tumour promoting potency (rosenkranz et al., 1997). recently cyanobacterial extracts and exudates were determined to be potent in vitro inhibitors of gjic (bláha et al., 2010; novakova et al., 2011). inhibition of gjic by these extracts were independent of the well-recognized tumour promoting cyanotoxins, mcs or cylindrospermopsin, indicating the existence of not-yet-identified toxic compounds (novakova et al., 2013). aquatic contaminants, such as polycyclic aromatic hydrocarbons (pahs) or polychlorinated biphenyls synergistically potentiated the inhibitory effects of cyanobacterial extracts on gjic (novakova et al., 2012), which further highlights the need to bioassay these mixtures for adverse effects in the complex assessment of drinking water quality, and efficacy of treatment technologies. in fact, in vitro assessment of gjic has been successfully used along with chemical analysis as a principal bioassay to study tumour promoting activity of water chlorination by-products (hakulinen et al., 2004; nishikawa et al., 2006). similarly, gjic assay has been applied to evaluate the efficiency of removing different anthropogenic contaminants and their toxic by-products with ozone, such as polycyclic aromatic hydrocarbons (upham et al., 1994; upham et al., 1995; herner et al., 2001; luster-teasley et al., 2002; luster-teasley et al., 2005) and various pesticides (upham et al., 1997; masten et al., 2001; wu et al., 2007). the in vitro bioassays used in these studies to assess gjic were based on the scrapeload dye transfer (sl-dt) technique (el-fouly et al., 1987). this sl-dt assay provides fast (minutes of exposure) and integrative responses, which reflect dysregulations of different cell processes and multiple signalling pathways controlling gjic (upham et al., 2016). dysregulation of gjic is an epigenetic, phenotypic marker for determining tumour promotional activity, which contrasts and compliments the more commonly used genotoxic and specific nuclear receptor transactivation assays (leusch and snyder, 2015). gjic can be assessed in vitro using diverse non-tumorigenic cells, with a rat liver epithelial cell line, wb-f344, being one of the most widely used gjic model for the assessment of tumour promoting activity, as well as determining chemopreventive effects of chemicals (upham et al., 1998; sovadinova et al., 2015; babica et al., 2016b). to further validate tumorigenic activity, compounds that dysregulate gjic are often tested for effects on signal transduction pathways implicated in neoplastic transformation, such as mitogen-activated protein kinases (mapk-erk1/2 and mapk-p38) (upham et al., 2008; osgood et al., 2014; babica et al., 2016a). our objective was to evaluate and compare the effino nco mm er cia l u se on ly chlorination and ozonation of cyanobacterial extract 109 cacy of two broadly used drinking water treatment oxidation technologies, namely chlorination and ozonation, on the removal of known cyanotoxin mc concentrations as well as on changes in biological effects that are independent of the mc content (i.e., removal of overall cytotoxicity and tumour promotional potency). we used a natural bloom sample that was dominated by the cosmopolitan and environmentally relevant bloom-forming cyanobacteria m. aeruginosa. extracts were prepared and characterized for content of mc, total organic carbon (toc) concentration, initial cytotoxicity, effect on gjic and modulation of signalling kinases (mapks). these cyanobacterial extracts were treated by chlorine or ozone, and evaluated for the changes in the toxin content, toc, and in vitro cytotoxicity and tumour promoting activity. methods cyanobacterial sample the sample of toxic cyanobacterial water bloom was collected from a lake located within the campus of michigan state university (east lansing, mi, usa; 42°40’50.09”n, 84°29’14.27”w) in september 2008 using a 20 μm plankton net. the bloom was dominated by microcystis species: m. aeruginosa (>50% of the cell counts) accompanied by m. flos-aquae (~20%) and m. ichthyoblabe (~20%). the biomass was freeze-dried and 38 g of dry weight (dw) was extracted by 20 min sonication (fisher sonic dismembrator model 300; fisher scientific, pittsburgh, pa, usa) while stirring on ice with 566 ml of 50% methanol (i.e., 66.7 g dw l–1 equivalent). the samples were centrifuged at 31,000 × g and the supernatant fraction was collected and dried using a vacuum evaporator. the dry extract was dissolved in 47.2 ml of milliq water (milliq synthesis a10; millipore, billerica, ma, usa) to obtain the final concentrated extract corresponding to 800 g dw of original biomass per one litter of water. chlorination and ozonation the extract was aliquoted into 4 ml fractions to be treated by chlorination or ozonation as summarized in tab. 1. chlorination was carried out with sodium hypochlorite (naocl) in a phosphate buffer (0.5 m k2hpo4-0.293 m naoh, ph 7.0±0.2) according to the method 5710c (clesceri et al., 1998). the sample was treated for 30 min to 24 h with naocl at different concentrations of free chlorine (7 to 1000 mg l–1) corresponding to contact time (ct) values 0.21×103, 7×103, 50×103 and 1440×103 mg min l–1. the oxidation reaction was stopped by addition of 10% w/v of nahso3. vehicle controls were prepared from milliq water equally treated with chlorine and quenched. no residual chlorine was present in the vehicle controls after the quenching. to increase the weight ratio of chlorine to dry weight of the extract (or toc or mc concentration), the concentration of the original biomass was also diluted fourand eight times (i.e., “¼” or “⅛”) before the chlorination step with 500 mg l–1 of chlorine for 100 min. the original extract dissolved in milliq water was ozonated for 30 min using a commercial ozone generator (oz2pcs-v; ozotech, yreka, ca, usa) at the concentration of o3 (gas)=5.0 mg l–1 with a gas flow rate of one l min–1 and temperature of 20.0±0.5°c to maximize ozone dissolution. microcystin analysis the concentrations of mcs in the original biomass as well as in the samples after the chlorination and ozonation treatment were analysed by hplc-uv/dad following the procedure described earlier (babica et al., 2006). toxins were identified based on their retention times and characteristic uv absorption spectra and quantified using the calibration curves of standards of mc-rr, -yr, -lr, -lw, -lf, and nodularin. an example of the hplc chromatograph of the cyanobacterial extract recorded at 238 nm is in supplementary fig. 1. toc analysis toc of the extract before and after chlorination and ozonation was determined using liquitoc analyzer (elementar analysensysteme, hanau, germany) where measurements were made by high temperature oxidation of the carbon (850-900°c) and detection of co2 by an ndir photometer. bioassays wb-f344 rat liver non-tumorigenic stem-like cells (tsao et al., 1984) were cultured in d-media (kao et al., 1997) with 5% v/v of fetal bovine serum (gibco, life technologies, grand island, ny, usa) at 37°c and 5% co2. cells were cultured to full confluence for 48 h in 35 mm tissue culture dishes (costar; cambridge, ma, usa). these confluent cells were used for the various time and dose related experiments. the vehicle controls were water chlorinated or ozonated using the same conditions as applied for studied samples. a sample or vehicle was added directly to the cell culture medium in the dish and gently mixed. non-treated cells were used as negative controls. the final concentrations of extracts in the bioassays were expressed as the original weight of dry biomass used for extract preparation per unit volume (g dw l–1). viability/cytotoxicity was tested after 30-min and 24-h exposures using the neutral red assay as reported before (babica et al., 2016b). the method determines viable cells capable of neutral red inclusion into lysosomes (borenfreund and puerner, 1985). viability was expressed as the fraction of negative (non-treated) control (foc). no nco mm er cia l u se on ly i. sovadinová et al.110 ta b. 1 .c on di tio ns o f c hl or in at io n an d oz on at io n of th e st ud ie d cy an ob ac te ri al e xt ra ct . a bb re vi at io n t re at m en t t yp e d os e of c l o r o 3 t re at m en t d ur at io n to ta l m c m c -l r t o c 2 4h c el l v ia bi lit y 3 0m in 3 0m in 3 0m in a nd c t v al ue (m g l –1 ) (µ g g– 1 d w )° ( µ g g– 1 d w )° (µ g g– 1 d w )° i c 50 (9 5% c i) g ji c e r k 1/ 2 p 38 ( m g m in l –1 ) l o e c i c 50 (9 5% c i) l o e c l o e c (g d w l –1 ) l o e c (g d w l –1 ) ( g d w l –1 ) ( g d w l –1 ) n t no tr ea tm en t - - 51 7 4 11 1 04 ,7 61 1 1. 6# 7. 8# ( 0. 52 -1 2. 41 )§ (0 .4 19. 86 )§ ( 10 525 14 )§ ( 10 .2 -1 3. 1) ( 7. 0 to 8 .7 ) 1 2 8 8 4 c l( 730 ) c hl or in at io n 7 3 0 m in 5 59 4 43 1 14 ,7 61 1 4. 8 10 .0 0 .2 1× 10 3 (0 .5 613 .4 2) (0 .4 410 .6 3) ( 11 427 39 ) (1 2. 617 .3 ) (9 .3 -1 0. 7) 1 2 8 8 4 c l( 70 -1 00 ) c hl or in at io n 70 10 0 m in 44 7 3 52 88 ,3 55 1 4. 3 1 0. 7* ** ^ 7 × 10 3 (0 .4 510 .7 4) (0 .3 58. 44 ) (8 821 21 ) ( 12 .2 -1 6. 5) (9 .7 -1 1. 8) 1 2 8 1 6 4 c l( 50 010 0) c hl or in at io n 5 00 1 00 m in 46 7 3 71 86 ,8 95 9. 1 1 2. 6* ** 5 0× 10 3 (0 .4 711 .2 0) (0 .3 78. 91 ) (8 720 85 ) (6 .9 -1 2. 2) (1 1. 214 .1 ) 1 2 8 1 6 4 ¼ c l( 50 010 0) $ c hl or in at io n 5 00 1 00 m in 34 6 2 75 77 ,2 29 1 0. 4 9 .6 5 0× 10 3 (0 .3 58. 31 ) (0 .2 86. 60 ) (7 718 53 ) (7 .1 -1 5. 4) (8 .4 -1 1. 0) 1 2 8 8 4 ⅛ c l( 50 010 0) $ c hl or in at io n 5 00 1 00 m in 41 2 3 31 81 ,1 93 7. 0 8 .1 5 0× 10 3 (0 .4 19. 88 ) (0 .3 37. 94 ) (8 119 49 ) (7 .0 -1 5. 5) ( 6. 79. 7) 1 2 8 n. a. n .a . c l( 10 00 -2 4h ) c hl or in at io n 1 00 0 24 h 3 67 2 95 78 ,7 54 1 2. 3 1 2. 7* ** 1 44 0× 10 3 ( 0. 37 -8 .8 1) ( 0. 29 -7 .0 7) (7 918 90 ) (9 .7 -1 5. 7) (1 0. 715 .0 ) 1 2 12 16 4 o 3 o zo na tio n 5 °° 3 0 m in n .d . n .d . 7 0, 62 2 > 24 > 24 (7 116 95 ) > 24 2 4 > 16 4 c t v al ue s: th e co nc en tr at io n of fr ee c hl or in e m ul tip lie d by th e co nt ac t t im e w ith th e sa m pl e be in g di si nf ec te d; it is e xp re ss ed in u ni ts o f m g m in l –1 . ° t he c on ce nt ra tio n is c al cu la te d pe r g d w o f o ri gi na l bi om as s be fo re m et ha no l e xt ra ct io n. # p la in te xt r ep re se nt s ic 50 va lu es w ith 9 5% c on fi de nc e in te rv al s in p ar en th es es , i ta lic iz ed n um be rs s ho w l o e c v al ue s. § v al ue s in p ar en th es es in di ca te th e ra ng e of ac tu al m c a nd t o c c on ce nt ra tio ns (m g l –1 ) i n th e ex pe ri m en ts (c al cu la te d fo r t he ra ng e of te st ed c on ce nt ra tio ns o f 1 -2 4 g d w l –1 ). ^ s ig ni fi ca nt ly d if fe re nt th an ic 50 va lu e of n t e xt ra ct (s tu de nt ’s tte st , p< 0. 05 ). $ b ef or e ch lo ri na tio n, th e co nc en tr at io n of o ri gi na l b io m as s w as d ilu te d fo ur (¼ ) o r e ig ht ti m es (⅛ ). °° g as fl ow ra te o f 1 l m in –1 fo r 3 0 m in ; n. d. , n ot d et ec te d: m in im al d et ec tio n lim it (m d l ) f or an in di vi du al m c v ar ia nt =1 .3 µ g g– 1 d w ; n .a ., no t a na ly se d. no nco mm er cia l u se on ly chlorination and ozonation of cyanobacterial extract 111 tumour promotion assay based on gjic assessment used only the concentrations that were determined to be noncytotoxic using the neutral red assay. cells were exposed for 30 min to the tested samples or corresponding vehicle. treatment with 1-methylanthracene (70 µm, 30 min) was used as a positive control inducing complete inhibition of gjic. gjic was assessed using modified sldt technique (el-fouly et al., 1987; babica et al., 2016c; upham et al., 2016) . the migration of the dye through gap junctions was visualized with a nikon eclipse te3000 phase contrast/fluorescent microscope and the images digitally captured with nikon ez cool snap ccd camera (nikon instruments, melville, ny, usa), where three representative images were acquired from each dish. the area of dye transfer was measured for each image using imagej (https://imagej.nih.gov/ij/). the measured areas were adjusted by subtracting an area of the dye transfer in the positive control with completely inhibited gjic. adjusted areas from each image were compared with an averaged adjusted area of the negative control and expressed as foc. activation of regulatory kinases mapk erk1/2 and p38 after a 30 min exposure to the sample was determined by western blotting. western blot analyses were done as reported previously (babica et al., 2016a). briefly, the proteins were extracted with 20% sds solution containing inhibitors of proteases and phosphatases, and the protein concentration of the cell lysates was determined with dc assay kit (bio-rad, hercules, ca, usa). the proteins (20 µg per lane) were separated on 12.5% sds-page (laemmli, 1970) and then electrophoretically transferred to a 0.45 µm pvdf membrane (millipore). to visualize activated, i.e. phosphorylated erks and p38, we used rabbit phospho-specific polyclonal antibodies directed to erk-1 phosphorylated at thr 202/tyr204, and erk-2 phosphorylated at thr185/tyr187 (cell signaling #9101s, danvers, ma, usa) and directed to p38 phosporylated at thr180/tyr182 (zymed #36-8500, san francisco, ca, usa), and secondary donkey anti-rabbit igg conjugated with horse radish peroxidase (amersham bioscience # na934v, life science, denver, co, usa). levels of glyceraldehyde 3-phosphate dehydrogenase (gapdh) was used as a housekeeping protein, and determined with mouse anti-gapdh antibodies (chemicon #mab374, millipore) and secondary sheep anti-mouse igg conjugated with horse radish peroxidase (amersham bioscience #na931v; life science). the erk, p38 and gapdh protein bands were detected using the ecl supersignal west pico chemiluminesence detection kit (pierce, arlington heights, il, usa) and bio-rad image analyzer. data analyses and statistics at least three independent experiments were done for each treatment (except for western blot analyses), and the mean ± standard deviations from independent experiments were calculated. ic50 concentrations causing 50% inhibition of the studied effects and their 95% confidence intervals were derived using non-linear regression in graphpad (graphpad software inc., la jolla, ca, usa). one-way analysis of variance (anova) followed by dunnett’s post-hoc test was used to identify treatments significantly different from the control. differences between ic50 values were assessed by student’s t-test. p-values less than 0.05 were considered statistically significant. results the original biomass contained 517 μg of total mc per g of dw with mc-lr being the dominant variant (411 μg g–1 dw, 79%, tab. 1). the other variants present included mc-rr (2%), mc-lw (2%), mc-lf (3%) and two other structurally unidentified mc variants (14%). the total mc concentration in the original extract (corresponding to 800 g biomass dw l–1) was 414 mg l–1 (mclr concentration: 329 mg l–1). toc concentration in the non-treated extract of 800 g biomass dw l–1 was 83.81 g l–1, which represents 105 mg of extractable total organic carbon per g of biomass dw. ozonation of cyanobacterial samples had direct effects on the concentrations of organic matter in the biomass. as shown in tab. 1, the 30-min treatment with ozone caused complete degradation of mcs and 33% reduction of toc compared to non-treated extract. the chlorination with low dose of chlorine (7 mg l–1) for short time (30 min) had no effect on mc and toc concentrations. the higher doses (70 to 1000 mg l–1) of chlorine and longer treatments (100 min to 24 h) were more effective in reducing mc levels (by 10 to 33%), and this reduction was dependent on chlorine concentrations and treatment times. there was no observed shift in the proportion of individual mc congeners with the mc-lr being the dominant variant. toc degradation by chlorine (by 16 to 25%) correlated well with mc degradation. the increase of the weight ratio of chlorine to original biomass by diluting nontreated extract before chlorination (tab. 1, variants ¼ or ⅛ cl(500-100)) slightly improved reduction of mc and toc amount, when the results were adjusted by the corresponding dilution factor and then compared to the nondiluted extract chlorinated under the same conditions (cl(500-100)). the original extract (non-treated, nt) showed no significant cytotoxic effects on wb-f344 during 30-min exposure within the range of concentrations tested (up to 24 g dw l–1; negative data not shown). after longer exposure time (24 h), the sample significantly decreased viability of wb-f344 cells (fig. 1a) with the calculated ic50=11.6 g dw l–1 (tab. 1). the lowest observed effect concentration (loec) inducing statistically significant reduction of the cell viability was 12 g dw l–1 (fig. 1a). no nco mm er cia l u se on ly i. sovadinová et al.112 the effects of the nt extract on gjic in wb-f344 cells after 30-min exposure are shown in fig. 2a. the sample had pronounced inhibitory effects on gjic, which were concentration-dependent with the calculated ic50=7.8 g dw l–1 (tab. 1). loec concentration of nt extract for inhibition of intercellular communication was 8 g dw l–1 (containing 4.1 mg l–1 mcs) (fig. 2a). the tumour promotional activity of the nt sample indicated by its effect on gjic was also confirmed by additional experiments assessing phosphorylation of signalling protein fig. 1. the effect of chlorination (b-e) and ozonation (f) of the tested cyanobacterial extract on viability in wb-f344 cells after 24-h exposure to 6 different concentrations (4-24 g dw l–1). data are fractions of controls (foc) as means ± standard deviations of independent repetitions of the experiment (n≥3). significant differences from the vehicle control (a) are indicated by asterisks (one-way anova followed by dunnet’s post-hoc test; *p≤0.05; **p≤0.01; ***p≤0.001; ****p≤0.0001). nt, original non-treated extract; cl, chlorinated extracts (in parentheses free cl concentrations 7, 70, 500 and 1000 mg l–1; treatment durations 30 min, 100 min, 24 h); o3, ozonated extract (30 min, the flow rate of 1 l min–1 with 5 g o3 m–3). no nco mm er cia l u se on ly chlorination and ozonation of cyanobacterial extract 113 kinases in wb-f344 cells (mapk-erk 1/2 and p38, fig. 3a). rapid and clear concentration-dependent activation of both mapks was observed after 30-min exposure. the loec values for hyperphosphorylation of mapkerk 1/2 and p38 were 8 or 4 g dw l–1 of the nt extract, respectively, with maximal effects observed at the highest tested concentration of 16 g dw l–1 (fig. 3a). ozonation and chlorination of the original sample had a pronounced effect on the biological activity. the cytotoxic effect was completely eliminated by the ozonation of the extract (fig. 1f). at the same time, none of the chlorination experimental protocols had any significant effect on cytotoxicity of the original nt extract (figs. 1be), with the estimated ic50 values ranging between 9.1 and 14.8 g dw l–1 and not being significantly different (p<0.05) from the ic50 value for the nt extract (tab. 1). also, the loec values for viability of cells exposed to chlorinated extracts remained at 12 g dw l–1. interestingly, slight but non-significant increase of ic50 values was associated with lower rather than higher chlorine doses. the increase of weight ratio of chlorine to the original biomass by diluting non-treated extract before chlorination did not decrease the cytotoxicity with the ic50 values of 10.4 g dw l–1 for ¼ cl(500-100) and 7.0 g dw l–1 for ⅛ cl(500-100) (tab. 1, supplementary fig. 2). the corresponding vehicle controls for these chlorination conditions caused 20 to 30% significant decrease in wbf344 cell viability when compared to non-treated control (supplementary fig. 2). the ozone application was also highly potent in eliminating tumour promotional activity of the original nt extract, as shown in fig. 2f. after 30-min ozonation, the loec value for gjic inhibition was 24 g dw l–1, i.e. three times higher than in nt extract. gjic was reduced only by 30% at 24 g dw l–1 concentration (fig. 2f), while nearly 100% inhibition of gjic was observed at 16 to 24 g dw l–1 of nt extract (fig. 2a). in contrast, the chlorination treatments had much less pronounced effects on the inhibition of gjic, with the ic50 values of the chlorinated extracts ranging between 10.0-12.7 g dw l–1. although these ic50 values were relatively similar to the ic50 estimated for nt extract, they were significantly higher except for the lowest free chlorine doses and shortest treatment (p<0.05), and their increase depended on the chlorine doseand treatment duration (tab. 1). when the weight-ratio of chlorine to the original biomass was increased by diluting the non-treated extract before chlorination (supplementary fig. 3), there was no observable decrease in the dysregulation of gjic induced as compared to the original nt extract, with the ic50 values after chlorination being 8.1 to 9.6 g dw l–1 (tab. 1, supplementary fig. 2). both ozone and chlorine treatments significantly attenuated the activation of mapk erk1/2 induced by nt extract (fig. 3). ozonation was more effective, when no activations of the erk1/2 kinase were observed at concentrations between 4 to 16 g dw l–1. chlorination also resulted in reduced levels of mapk erk1/2 activation (fig. 3, tab. 1). the protective effects had biphasic character, when initially increased but then became less apparent with an increase in chlorine dose and chlorination time (fig. 3, tab. 1). the effects of both ozone and chlorine treatments on the activation of p38 mapk were much less pronounced when compared to mapk erk1/2, and activation of p38 was still apparent for the ozonated extract as well as most of the chlorinated extracts even at the lowest experimental concentration of 4 g dw l–1 (fig. 3). discussion adverse effects of toxic cyanobacteria on human health remain a major issue for both researchers and water managers. our study confirmed that rapid inhibitions of gjic and activations of mapk-erk1/2 might be a common effect induced by bloom-forming cyanobacteria. the effective concentration for 30-min gjic inhibition (ic50=8 g dw l–1) was similar to previously reported values for other extracts from natural blooms dominated by microcystis sp. (ic50=4 or 6 g dw l–1, respectively) (bláha et al., 2010; novakova et al., 2011). modulation of these cellular events by chemicals in vitro are considered to be relevant biomarkers of tumour promoting potency in vivo, as was demonstrated e.g. for tumour promoting phorbol esters like tpa (madhukar et al., 1996), organochlorine pesticides (trosko et al., 1987), pcbs (kang et al., 1996), low molecular weight pahs (bláha et al., 2002), clofibrate, phenobarbital, perfluorooctanoic acid, or organic peroxides (upham et al., 2007; upham et al., 2009; vinken et al., 2009). mcs and other cyanotoxins like cylindrospermopsin have been shown to modulate development of tumours (nishiwaki-matsushima et al., 1992; falconer and humpage, 2005; svircev et al., 2010; zegura et al., 2011; de la cruz et al., 2013). however, neither mc-lr nor cylindrospermopsin had any direct effect on rapid inhibition of gjic or activation of mapk erk 1/2 indicating that other metabolites in cyanobacteria might be responsible for their gjic-dependent tumour promoting activity (bláha et al., 2010; novakova et al., 2011). however, these metabolites have not been identified yet. in addition, our present study demonstrates, for the first time, a rapid activation of another mapk-p38 by cyanobacterial environmental extract in the cell line wbf344, which possesses characteristics of liver progenitor cells (babica et al., 2016a, b). mapk-p38 is a critical participant in cellular stress responses and has a key role in inflammation, as well as in tissue homeostasis, by controlling cell proliferation, differentiation, death, survival and the no nco mm er cia l u se on ly i. sovadinová et al.114 migration of specific cell types (didonato et al., 2012). in contrast to mapk-erk1/2, which are activated by mitogens or growth factors, mapk-p38 is activated by environmental and genotoxic stresses including hypoxia, uv, ros, hyperosmolarity, and heat shock, and its activation has been linked to protein phosphatase 2 (pp2a) (nebreda and porras, 2000; wagner and nebreda, 2009). indeed, mc-lr, a known pp2a inhibitor, has been previously reported to activate mapks erk1/2, p38 or jnk in rodent liver in vivo and also in experiments with various cell lines fig. 2. the effect of chlorination (b-e and ozonation (f) of the tested cyanobacterial extract on gap-junctional intercellular communication (gjic) in wb-f344 cells after 30-min exposure to 6 different concentrations (1-24 g dw l–1). data are fractions of controls (foc) as means ± standard deviations of independent repetitions of the experiment (n≥3). significant differences from the nt extract (a) are indicated by asterisks (one-way anova followed by dunnet’s post hoc test; *p≤0.05; **p≤0.01; ***p≤0.001; ****p≤0.0001). nt, original non-treated extract; cl, chlorinated extracts (in parentheses free cl concentrations 7, 70, 500 and 1000 mg l–1; treatment durations 30 min, 100 min, 24 h); o3, ozonated extract (30 min, the flow rate of 1 l min–1 with 5 g o3 m–3). no nco mm er cia l u se on ly chlorination and ozonation of cyanobacterial extract 115 in vitro. however, mc-lr effects on mapks seem to be dependent on the kinase type, a cell type, and probably also exposure times and concentrations. depending on the study, mc-lr was found to activate erk1/2 but not p38 (dias et al., 2010; zhang et al., 2013; adamovsky et al., 2015), p38 but not erk1/2 (meng et al., 2011; lezcano et al., 2012), and both erk1/2 and p38 (komatsu et al., 2007; daily et al., 2010; sun et al., 2011; chen et al., 2012; liu et al., 2016; wang et al., 2017). the effective concentrations of mc-lr in these studies were typically between 1 and 10 µm (~1-10 mg l –1). in the present study, cytotoxicity, gjic and mapks were affected by the cyanobacterial extract diluted to 412 g dw l–1 containing mcs with concentrations between 2 and 6 mg l–1, which corresponds to 2 to 6 µm range and is quite comparable with the other in vitro studies reporting mapk activation by mcs. however, we previously demonstrated that tumour promoting events, such as rapid gjic inhibition and mapk erk1/2 activation in rat liver progenitor cells, are induced by other cyanobacterial metabolites but not mc-lr or cylindrospermopsin (bláha et al., 2010). our results suggest that mapk p38 can also be activated by transformation or degradation products of mcs, other compounds of cyanobacterial origin and/or their transformation or degradation products, since p38 was activated not only by mccontaining non-treated cyanobacterial extract, but also by ozonated extract without detectable levels of mcs. these findings suggest that progenitor cells, in comparison with differentiated hepatocytes, might be less prone to the effects of mcs, possibly due to limited expression of key proteins involved in mcs uptake and metabolism, such as organic-anion transporting polypeptides (oatps). nevertheless, progenitor cells can be apparently a target of other compounds present in cyanobacterial biomass, which are capable to induce toxic and tumour promoting effects in this specific population of liver cells (bláha et al., 2010) known to play a critical role in the maintenance of liver tissue homeostasis, liver regeneration and hepatocarcinogenesis (canovas-jorda et al., 2014). the evidence supporting the existence of other components of cyanobacterial biomass contributing to the tumour promoting and toxic effects of complex cyanobacterial samples emphasize the need for effectbased evaluation of the efficacy of water treatment technologies in addition to chemical analyses. different physicochemical purification processes employed in dwtps may have different efficacies in removing the target contaminant, such as well recognized toxicants (such as mcs) vs. elimination of the overall toxicity. our results demonstrate that ozone effectively and rapidly removes the mc fraction of the complex cyanobacterial samples. although the highest chlorine dose resulted in a decrease of toc that was comparable to the ozone treatment, reactive chlorine was not as effective compared to ozone in removal of mcs. similarly, removal of cytotoxicity and overall epigenetic toxicity of the studied sample by ozonation appeared to be much more effective than chlorination, although ozone had a less pronounced effect in decreasing p38 activation as compared to erk1/2 and gjic. with regard to chlorine the literature demonstrate that its application removes mcs but the efficiency of removal decreases with increases in ph and dissolved organic material along with formation of less effective oxidant clo– (merel et al., 2010). several studies agreed that 0.5 mg of residual chlorine per liter should efficiently remove pure mcs in distilled water during 5-30 min (depending on mc concentration) at ph lower than 8 (nicholson et al., 1994; newcombe and nicholson, 2004; acero et al., 2005). similar scenarios are also expected at dwtp at environmentally relevant concentrations of mcs (merel et al., 2009). several previous studies using pp-inhibition as a biomarker of toxicity as well as other bioassays reported a decrease in mc concentration and toxicity after chlorination of cyanobacterial samples (nicholson et al., 1994; tsuji et al., 1997; rodriguez et al., 2008; merel et al., 2010). for example, the chlorination of m. aeruginosa extract (16 mg dw l–1) with the dose of 1 mg l–1 chlorine and contact time of 30 min (ct value=30) effectively removed 95% of mcs (initial concentration: 192 µg l–1) and completely eliminated the acute toxicity of this extract in a mouse bioassay (nicholson et al., 1994). interestingly, our study showed that chlorination of cyanobacterial extract was less effective. the dose of 7 mg of chlorine per liter for 30 min at ph 7.2 did not decrease mcs nor toc concentrations. higher doses and longer treatment times removed up to 1030% of the original mcs and toc levels, and caused only moderate reduction of toxicity (cytotoxicity, gjic inhibition and mapk activations). chlorination apparently removed some compounds responsible for the inhibition of gjic and activation of erk1/2 as reflected by the slight, yet statistically significant, increase in respective ic50 or loec values, and had only a minor effect on cytotoxicity and p38 activation. the lower effectiveness of chlorination observed in our study could be explained by the interactions of chlorine with the relatively higher concentrations of organic matter, which might have reduced the effectiveness of the oxidation process due to the competition between the toxins and the dissolved organic carbon reacting with the oxidant (rodriguez et al., 2008). an important problem associated with the chlorine application is the formation of by-products such as halogenated organic compounds, especially in the presence of high amounts of organic matter. these by-products can have toxic or potential carcinogenic potencies (neale et al., 2012); and also for mcs, chlorine was shown to cause substitutions and modifications of the toxic adda moiety no nco mm er cia l u se on ly i. sovadinová et al.116 fig. 3. activation of mitogen-activated protein kinases (mapks) by studied samples after 30-min exposure to 3 different concentrations (4-16 g dw l–1). phosphorylation of extracellular receptor kinases 1 and 2 (erk1/2) and p38 was determined by western blotting (a). the bar graphs (b) show values from the densitometric image analysis normalized to negative control (nc=1). nt, original non-treated extract; cl, chlorinated extracts (in parentheses free cl concentrations 7, 70, 500 and 1000 mg l–1; treatment durations 30 min, 100 min, 24 h), o3 – ozonated extract (30 min, the flow rate of 1 l min–1 with 5 g o3 m–3); nc, negative control (no treatment of the cells); pc, positive control for erk1/2 activation (12-o-tetradecanoyl phorbol-13-acetate, 10 nm, 30 min). no nco mm er cia l u se on ly chlorination and ozonation of cyanobacterial extract 117 (tsuji et al., 1997; merel et al., 2009). in addition, de novo formation of chlorinated by-products with potencies to affect gjic and activate intracellular signalling (hakulinen et al., 2004; nishikawa et al., 2006) should also be considered, and could be related to the weak efficiency in removal of gjic inhibitions and mapk activations during the chlorination as observed in the present study. possible toxicity of chlorinated by-products could also explain the observed biphasic effect, when lower doses of chlorine were slightly more effective in the elimination of cytotoxicity and mapks activation than the higher doses, although concentrations of toc and mcs were slightly but progressively reduced with increasing chlorine dose and treatment time. we demonstrated that ozonation completely removed mcs, substantial fractions of toc and protected against cytotoxicity, gjic inhibition or activation of erk1/2. these findings are in agreement with similar studies, which documented complete mc removal (5 mg l–1) by 2 mg l–1 o3 within 2 min (al momani and jarrah, 2010). further improvements in kinetics could be achieved by increased o3 doses and temperature, and decreased ph (al momani and jarrah, 2010; shawwa and smith, 2001). naturally, organic matter negatively reduces the efficiency of ozonation, but under realistic dwtp situations of levels as low as 0.05 mg l–1 of residual o3 assures mc removal (newcombe and nicholson, 2004; brooke et al., 2006). despite a high amount of organic material in our sample that also competes with toxins for ozone, ozonation was highly effective in mc reduction and elimination of toxicity even after short treatment. a high efficiency of oxidation of mc is known to be mediated by hydroxyl radicals attacking conjugated diene structure in mc followed by the cleavage of the adda side chain (responsible for ppase inhibition) and ultimately opening of the peptide ring (al momani and jarrah, 2010; miao et al., 2010). biological assessments using pp-inhibition assay or mouse test confirmed elimination of the toxicity along with the described structural changes of mc (brooke et al., 2006; miao et al., 2010). although ozone was quite efficient in removal of cytotoxic, gjic inhibiting and erk1/2 activating compounds in our study, it had only a partial effect on the removal of p38 activating components. this might indicate that p38 is not involved in gjic inhibition and its activation was caused by metabolites with different modes of actions (wagner and nebreda, 2009). with respect to the critical role of p38 in cellular responses to different types of stress and also in controlling the proliferation, differentiation, survival, migration and inflammatory responses of specific cell types, further research should address interactions of cyanobacterial metabolites with this signalling pathway and evaluate its relevance as a biomarker of environmental and genotoxic stress induced by cyanobacteria. interestingly, activation of p38 was found to be the most sensitive endpoint in this study, where the increased levels of p38 phosphorylation were observed after 30-min exposure to the non-treated extract at concentration of 4 g dw l–1, whereas significant inhibition of gjic and activation of erk1/2 occurred at concentrations 8 g dw l–1 and higher, and significant reduction of cell viability required 24-h exposure to 12 g dw l–1. inhibition of gjic and activation of mapks was induced by lower concentrations and after shorter exposures than the cytotoxic effects, which indicates that these cell signalling events were altered via rapid non-genotoxic and non-cytotoxic mechanisms. in vitro evaluation of gjic and mapks thus represent a simple and sensitive bioassay for assessment of ‘epigenetic toxicity’ and tumour promoting potential of complex cyanobacterial extracts, which is also suitable for effect-based studies focusing on the elimination of these hazardous properties of contaminated water. conclusions ozonation of an extract of a microcystis water bloom sample was shown to be a very effective method in the complete removal of mcs, as well as the substantial elimination of the overall cytotoxicity and tumour promotional potency. on the contrary, chlorination experiments, despite high doses and long exposures, were much less effective, and potentially led to the formation of by-products, which could add to the observed toxic effects. our study also demonstrated strong activations of p38 mapk by cyanobacterial samples, which were not effectively removed by chlorination and only partially by ozonation. with respect to the role of p38 in inflammation as well as maintenance of tissue homeostasis, further research should address interactions of cyanobacterial samples with this biomarker of cellular stress and evaluate its environmental relevance. in agreement with several recent reports, the study also demonstrates the need to enforce effect-based (bioassay) tools into the assessment of water quality and monitoring the efficacy of water treatment systems. acknowledgments supported by the czech science foundation project ga15-12408s; the czech republic ministry of education, youth and sports infrastructure projects no. lo1214 and lm2015051; long-term research development project rvo 67985939; and by national institute of environmental health sciences (niehs) grant #r01 es013268-01a2 to upham. we would like to thank dr. ondřej mikeš for the help with toc analysis. no nco mm er cia l u se on ly i. sovadinová et al.118 references acero jl, rodriguez e, meriluoto j, 2005. kinetics of reactions between chlorine and the cyanobacterial toxins microcystins. water res. 39:1628-1638. adamovsky o, moosova z, pekarova m, basu a, babica p, svihalkova sindlerova l, kubala l, blaha l, 2015. immunomodulatory potency of microcystin, an important water-polluting cyanobacterial toxin. environ. sci. technol. 49:12457-12464. al momani fa, jarrah n, 2010. treatment and kinetic study of cyanobacterial toxin by ozone. j. environ. sci. health a tox. hazard. subst. environ. eng. 45:719-731. babica p, kohoutek j, blaha l, adamovsky o, marsalek b, 2006. evaluation of extraction approaches linked to elisa and hplc for analyses of microcystin-lr, -rr and -yr in freshwater sediments with different organic material contents. anal. bioanal. chem. 385:1545-1551. babica p, zurabian r, kumar er, chopra r, mianecki mj, park js, jasa l, trosko je, upham bl, 2016a. methoxychlor and vinclozolin induce rapid changes in intercellular and intracellular signaling in liver progenitor cells. toxicol. sci. 153: 174-185. babica p, ctverackova l, lencesova z, trosko je, upham bl, 2016b. chemopreventive agents attenuate rapid inhibition of gap junctional intercellular communication induced by environmental toxicants. nutr. cancer 68: 827-837. babica p, sovadinova i, upham bl, 2016c. scrape loading / dye transfer, p. 133-144. in: m. vinken and s. johnstone (eds.), methods in molecular biology. 1437. gap junction protocols. springer. berry jp, gibbs pdl, schmale mc, saker ml, 2009. toxicity of cylindrospermopsin, and other apparent metabolites from cylindrospermopsis raciborskii and aphanizomenon ovalisporum, to the zebrafish (danio rerio) embryo. toxicon 53 289-299. bláha l, kapplová p, vondráček j, upham b, machala m, 2002. inhibition of gap-junctional intercellular communication by environmentally occurring polycyclic aromatic hydrocarbons. toxicol. sci. 6543-51. bláha l, babica p, maršálek b, 2009. toxins produced in cyanobacterial water blooms – toxicity and risks. interdiscip. toxicol. 2:36-41. bláha l, babica p, hilscherová k, upham bl, 2010. inhibition of gap-junctional intercellular communication and activation of mitogen-activated protein kinases by cyanobacterial extracts indications of novel tumor promoting cyanotoxins? toxicon 55:126-134. borenfreund e, puerner ja, 1985. toxicity determined in vitro by morphological alterations and neutral red absorption. toxicol. lett. 24:119-124. brooke s, newcombe g, nicholson b, klass g, 2006. decrease in toxicity of microcystins la and lr in drinking water by ozonation. toxicon 48:1054-1059. campos a, vasconcelos v, 2010. molecular mechanisms of microcystin toxicity in nnimal cells. int. j. mol. sci. 11:268-287. canovas-jorda d, louisse j, pistollato f, zagoura d, bremer s, 2014. regenerative toxicology: the role of stem cells in the development of chronic toxicities. expert opin. drug metab. toxicol. 10:39-50. chen dn, zeng j, wang f, zheng w, tu ww, zhao js, xu j, 2012. hyperphosphorylation of intermediate filament proteins is involved in microcystin-lr-induced toxicity in hl7702 cells. toxicol. lett. 214:192-199. clesceri ls, greenberg ae, eaton ad. 1998. standard methods for examinations of water and wastewater. 20th edition, american public health association, washington dc: 1496 pp. codd ga, lindsay j, young fm, morrison lf, metcalf js. 2005. cyanobacterial toxins, p. 1-23. in: j. huisman, h. c. p. matthijs and p. m. visser (eds.), harmful cyanobacteria. springer. daily a, monks nr, leggas m, moscow ja, 2010. abrogation of microcystin cytotoxicity by map kinase inhibitors and n-acetyl cysteine is confounded by oatpib1 uptake activity inhibition. toxicon 55:827-837. de la cruz aa, hiskia a, kaloudis t, chernoff n, hill d, antoniou mg, he x, loftin k, o’shea k, zhao c, pelaez m, han c, lynch tj, dionysiou dd, 2013. a review on cylindrospermopsin: the global occurrence, detection, toxicity and degradation of a potent cyanotoxin. environ. sci. proc. imp. 15:1979-2003. dias e, matos p, pereira p, batoreu mc, silva mj, jordan p, 2010. microcystin-lr activates the erk1/2 kinases and stimulates the proliferation of the monkey kidney-derived cell line vero-e6. toxicol. in vitro 24:1689-1695. didonato ja, mercurio f, karin m, 2012. nf-kappab and the link between inflammation and cancer. immunol. rev. 246:379-400. el-fouly mh, trosko je, chang c-c, 1987. scrape-loading and dye transfer. exp. cell res. 168:422-430. falconer ir, humpage ar, 2005. health risk assessment of cyanobacterial (blue-green algal) toxins in drinking water. int. j. environ. res. public health. 2:43-50. grosse y, baan r, straif k, secretan b, el ghissassi f, cogliano v, 2006. carcinogenicity of nitrate, nitrite, and cyanobacterial peptide toxins. lancet oncol. 7:628-629. hakulinen p, maki-paakkanen j, naarala j, kronberg l, komulainen h, 2004. potent inhibition of gap junctional intercellular communication by 3-chloro-4-(dichloromethyl)-5hydroxy-2(5h)-furanone (mx) in balb/c 3t3 cells. toxicol. lett. 151:439-449. harke mj, steffen mm, gobler cj, otten tg, wilhelm sw, wood sa, paerl hw, 2016. a review of the global ecology, genomics, and biogeography of the toxic cyanobacterium, microcystis spp. harmful algae 54:4-20. herner ha, trosko je, masten sj, 2001. the epigenetic toxicity of pyrene and related ozonation byproducts containing an aldehyde functional group. environ. sci. technol. 35: 3576-3583. hiskia a, dionysiou dd, antoniou m, kaloudis t, 2016. water treatment for purification from cyanobacteria and cyanotoxins. 1st ed. wiley-blackwell: 272 pp. ibelings bw, bormans m, fastner j, visser pm, 2016. cyanocost special issue on cyanobacterial blooms: synopsis a critical review of the management options for their prevention, control and mitigation. aquat. ecol. 50:595-605. kang ks, wilson mr, hayashi t, chang cc, trosko je, 1996. inhibition of gap junctional intercellular communication in normal human breast epithelial cells after treatment with pesticides, pcbs, and pbbs, alone or in mixtures. environ. health perspect. 104:192-200. kao cy, oakley cs, welsch cw, chang cc, 1997. growth requirements and neoplastic transformation of two types of normal human breast epithelial cells derived from reduction mammoplasty. in vitro cellular & developmental biology. animal 33:282-288. no nco mm er cia l u se on ly chlorination and ozonation of cyanobacterial extract 119 komatsu m, furukawa t, ikeda r, takumi s, nong q, aoyama k, akiyama s, keppler d, takeuchi t, 2007. involvement of mitogen-activated protein kinase signaling pathways in microcystin-lr-induced apoptosis after its selective uptake mediated by oatp1b1 and oatp1b3. toxicol. sci. 97: 407-416. kuiper-goodman t, falconer i, fitzgerald j. 1999. human health aspects, p. 113-153. in: i. chorus and j. bartram (eds.), toxic cyanobacteria in water: a guide to their public health consequences, monitoring and management. world health organization, geneva, switzerland. laemmli uk, 1970. cleavage of structural proteins during the assembly of the head of bacteriophage t4. nature 227:680-685. leusch fdl, snyder sa, 2015. bioanalytical tools: half a century of application for potable reuse. environmental science: water res. technol. 1:606-621. lezcano n, sedan d, lucotti i, giannuzzi l, vittone l, andrinolo d, mundina-weilenmann c, 2012. subchronic microcystin-lr exposure increased hepatic apoptosis and induced compensatory mechanisms in mice. j. biochem. mol. toxic. 26:131-138. liu j, wang b, huang p, wang h, xu k, wang x, xu l, guo z, 2016. microcystin-lr promotes cell proliferation in the mice liver by activating akt and p38/erk/jnk cascades. chemosphere 163:14-21. luster-teasley sl, yao jj, herner hh, trosko je, masten sj, 2002. ozonation of chrysene: evaluation of byproduct mixtures and identification of toxic constituent. environ. sci. technol. 36:869-876. luster-teasley sl, ganey pe, diorio m, ward js, maleczka re, jr., trosko je, masten sj, 2005. effect of byproducts from the ozonation of pyrene: biphenyl-2,2’,6,6’-tetracarbaldehyde and biphenyl-2,2’,6,6’-tetracarboxylic acid on gap junction intercellular communication and neutrophil function. environ. toxicol. chem. 24:733-740. madhukar bv, de feijter rupp hl, trosko je, 1996. pulse treatment with the tumor promoter tpa delays the onset of desensitization response and prolongs the inhibitory effect on gap junctional intercellular communication of a rat liver epithelial cell line wb f-344. cancer lett. 106:117-123. maier d, blaha l, giesy jp, henneberg a, kohler hr, kuch b, osterauer r, peschke k, richter d, scheurer m, triebskorn r, 2015. biological plausibility as a tool to associate analytical data for micropollutants and effect potentials in wastewater, surface water, and sediments with effects in fishes. water res. 72:127-144. masten sj, tian m, upham bl, trosko je, 2001. effect of selected pesticides and their ozonation by-products on gap junctional intercellular communication using rat liver epithelial cell lines. chemosphere 44:457-465. máthé c, beyer d, m-hamvas m, vasas g, 2016. the effects of microcystins (cyanobacterial heptapeptides) on the eukaryotic cytoskeletal system. mini rev. med. chem. 16:1063-1077. mcgorum bc, pirie rs, glendinning l, mclachlan g, metcalf js, banack sa, cox pa, codd ga, 2015. grazing livestock are exposed to terrestrial cyanobacteria. vet. res. 46:1-10. meng g, sun y, fu w, guo z, xu l, 2011. microcystin-lr induces cytoskeleton system reorganization through hyperphosphorylation of tau and hsp27 via pp2a inhibition and subsequent activation of the p38 mapk signaling pathway in neuroendocrine (pc12) cells. toxicology 290:218-229. merel s, lebot b, clément m, seux r, thomas o, 2009. ms identification of microcystin-lr chlorination by-products. chemosphere 74:832-839. merel s, clément m, thomas o, 2010. state of the art on cyanotoxins in water and their behaviour towards chlorine. toxicon 55:677-691. merel s, walker d, chicana r, snyder s, baurès e, thomas o, 2013. state of knowledge and concerns on cyanobacterial blooms and cyanotoxins. environ. int. 59:303-327. miao h-f, qin f, tao g-j, tao w-y, ruan w-q, 2010. detoxification and degradation of microcystin-lr and -rr by ozonation. chemosphere 79:355-361. moreira c, azevedo j, antunes a, vasconcelos v, 2013. cylindrospermopsin: occurrence, methods of detection and toxicology. j. appl. microbiol. 114:605-620. mowe mad, mitrovic sm, lim rp, furey a, yeo dcj, 2014. tropical cyanobacterial blooms: a review of prevalence, problem taxa, toxins and influencing environmental factors. j. limnol. 74:1005. neale pa, antony a, bartkow me, farre mj, heitz a, kristiana i, tang jy, escher bi, 2012. bioanalytical assessment of the formation of disinfection byproducts in a drinking water treatment plant. environ. sci. technol. 46:10317-10325. nebreda ar, porras a, 2000. p38 map kinases: beyond the stress response. trends biochem. sci. 25:257-260. newcombe g, nicholson b, 2004. water treatment options for dissolved cyanotoxins. j. water supply res. technol. 53:227-239. nicholson bc, rositano j, burch md, 1994. destruction of cyanobacterial peptide hepatotoxins by chlorine and chloramine. water res. 28:1297-1303. nishikawa a, sai k, okazaki k, son hy, kanki k, nakajima m, kinae n, nohmi t, trosko je, inoue t, hirose m, 2006. mx, a by-product of water chlorination, lacks in vivo genotoxicity in gpt delta mice but inhibits gap junctional intercellular communication in rat wb cells. environ. mol. mutagen. 47:48-55. nishiwaki-matsushima r, ohta t, nishiwaki s, suganuma m, kohyama k, ishikawa t, carmichael ww, fujiki h, 1992. liver tumor promotion by the cyanobacterial cyclic peptide toxin microcystin-lr. j. cancer res. clin. oncol. 118:420-424. novakova k, babica p, adamovsky o, blaha l, 2011. modulation of gap-junctional intercellular communication by a series of cyanobacterial samples from nature and laboratory cultures. toxicon 58:76-84. novakova k, blaha l, babica p, 2012. tumor promoting effects of cyanobacterial extracts are potentiated by anthropogenic contaminants-evidence from in vitro study. chemosphere 89:30-37. novakova k, kohoutek j, adamovsky o, brack w, krauss m, blaha l, 2013. novel metabolites in cyanobacterium cylindrospermopsis raciborskii with potencies to inhibit gap junctional intercellular communication. j. hazard. mater. 262:571-579. oberemm a, fastner j, steinberg cew, 1997. effects of microcystin-lr and cyanobacterial crude extracts on embryo-larval development of zebrafish (danio rerio). water res. 31:2918-2921. osgood rs, upham bl, hill t, helms kl, velmurugan k, babica p, bauer ak, 2014. polycyclic aromatic hydrocarbon-induced signaling events relevant to inflammation and tumorigenesis in lung cells are dependent on molecular structure. plos one 8:e65150. no nco mm er cia l u se on ly i. sovadinová et al.120 prasse c, stalter d, schulte-oehlmann u, oehlmann j, ternes ta, 2015. spoilt for choice: a critical review on the chemical and biological assessment of current wastewater treatment technologies. water res. 87:237-270. rodriguez em, acero jl, spoof l, meriluoto j, 2008. oxidation of mc-lr and -rr with chlorine and potassium permanganate: toxicity of the reaction products. water res. 42: 1744-1752. roegner af, brena b, gonzález-sapienza g, puschner b, 2014. microcystins in potable surface waters: toxic effects and removal strategies. j. appl. toxicol. 34:441-457. rosenkranz m, rosenkranz hs, klopman g, 1997. intercellular communication, tumor promotion and non-genotoxic carcinogenesis: relationships based upon structural considerations. mutat. res. 381:171-188. shawwa ar, smith dw, 2001. kinetics of microcystin-lr oxidation by ozone. ozone sci. eng. 23:161-170. sovadinova i, babica p, boke h, kumar e, wilke a, park js, trosko je, upham bl, 2015. phosphatidylcholine specific plc-induced dysregulation of gap junctions, a robust cellular response to environmental toxicants, and prevention by resveratrol in a rat liver cell model. plos one 10:e0124454. sun y, meng gm, guo zl, xu lh, 2011. regulation of heat shock protein 27 phosphorylation during microcystin-lrinduced cytoskeletal reorganization in a human liver cell line. toxicol. lett. 207:270-277. svircev z, krstic s, miladinov-mikov m, baltic v, vidovic m, 2009. freshwater cyanobacterial blooms and primary liver cancer epidemiological studies in serbia. j. environ. sci. health c environ. carcinog. ecotoxicol. rev. 27:36-55. svircev z, baltic v, gantar m, jukovic m, stojanovic d, baltic m, 2010. molecular aspects of microcystin-induced hepatotoxicity and hepatocarcinogenesis. j. environ. sci. health c environ. carcinog. ecotoxicol. rev. 28:39-59. trosko je, jone c, chang cc, 1987. inhibition of gap junctional-mediated intercellular communication in vitro by aldrin, dieldrin and toxaphene: a possible cellular mechanism for their tumor-promoting and neurotoxic effects. mol. toxicol. 1:83-94. tsao ms, smith jd, nelson kg, grisham jw, 1984. a diploid epithelial cell line from normal adult rat liver with phenotypic properties of ‘oval’ cells. exp. cell res. 154:38-52. tsuji k, watanuki t, kondo f, watanabe mf, nakazawa h, suzuki m, uchida h, harada k-i, 1997. stability of microcystins from cyanobacteria iv. effect of chlorination on decomposition. toxicon 35:1033-1041. upham b, masten sj, lockwood br, trosko je, 1994. nongenotoxic effects of polycyclic aromatic hydrocarbons and their ozonation by-products on the intercellular communication of rat liver epithelial cells. fundam. appl. toxicol. 23 470-475. upham bl, yao jj, trosko je, masten sj, 1995. determination of the efficacy of ozone treatment systems using a gap junction intercellular communication bioassay. environ. sci. technol. 29:2923-2928. upham bl, boddy b, xing x, trosko je, masten sj, 1997. non-genotoxic effects of selected pesticides and their disinfection by-products on gap junctional intercellular communication. ozone: sci. eng. 19:351-369. upham bl, weis lm, trosko je, 1998. modulated gap junctional intercellular communication as a biomarker of pah epigenetic toxicity: structure-function relationship. environ. health perspect. 106(suppl 4):975-981. upham bl, guzvic m, scott j, carbone jm, blaha l, coe c, li ll, rummel am, trosko je, 2007. inhibition of gap junctional intercellular communication and activation of mitogenactivated protein kinase by tumor-promoting organic peroxides and protection by resveratrol. nutr. cancer 57:38-47. upham bl, blaha l, babica p, park js, sovadinova i, pudrith c, rummel am, weis lm, sai k, tithof pk, guzvic m, vondracek j, machala m, trosko je, 2008. tumor promoting properties of a cigarette smoke prevalent polycyclic aromatic hydrocarbon as indicated by the inhibition of gap junctional intercellular communication via phosphatidylcholine-specific phospholipase c. cancer sci. 99:696-705. upham bl, park js, babica p, sovadinova i, rummel am, trosko je, hirose a, hasegawa r, kanno j, sai k, 2009. structure-activity-dependent regulation of cell communication by perfluorinated fatty acids using in vivo and in vitro model systems. environ. health perspect. 117:545-551. upham bl, sovadinova i, babica p, 2016. gap junctional intercellular communication: a functional biomarker to assess adverse effects of toxicants and toxins, and health benefits of natural products. j. vis. exp. 118:e54281. vinken m, doktorova t, decrock e, leybaert l, vanhaecke t, rogiers v, 2009. gap junctional intercellular communication as a target for liver toxicity and carcinogenicity. crit. rev. biochem. mol. biol. 44:201-222. wagner ef, nebreda ar, 2009. signal integration by jnk and p38 mapk pathways in cancer development. nat. rev. cancer 9:537-549. wang b, liu j, huang p, xu k, wang h, wang x, guo z, xu l, 2017. protein phosphatase 2a inhibition and subsequent cytoskeleton reorganization contributes to cell migration caused by microcystin-lr in human laryngeal epithelial cells (hep-2). environ. toxicol. 32:890-903. wernersson a, maggi c, carere m, 2014. technical report on aquatic effect-based monitoring tools. office for official publications of the european communities. westrick ja, szlag dc, southwell bj, sinclair j, 2010. a review of cyanobacteria and cyanotoxins removal/inactivation in drinking water treatment. anal. bioanal. chem. 397:705-1714. wiegand c, pflugmacher s, 2005. ecotoxicological effects of selected cyanobacterial secondary metabolites a short review. toxicol. appl. pharmacol. 203:201-218. wu j, lin l, luan t, chan gilbert ys, lan c, 2007. effects of organophosphorus pesticides and their ozonation byproducts on gap junctional intercellular communication in rat liver cell line. food chem.toxicol. 45:2057-2063. yu sz, 1995. primary prevention of hepatocellular carcinoma. j. gastroenterol. hepatol. 10:674-682. zanchett g, oliveira-filho e, 2013. cyanobacteria and cyanotoxins: from impacts on aquatic ecosystems and human health to anticarcinogenic effects. toxins 5:1896. zegura b, straser a, filipic m, 2011. genotoxicity and potential carcinogenicity of cyanobacterial toxins a review. mutat. res. 727:16-41. zhang j, chen j, xia z, 2013. microcystin-lr exhibits immunomodulatory role in mouse primary hepatocytes through activation of the nf-kappab and mapk signaling pathways. toxicol. sci. 136:86-96. zhou l, yu h, chen k, 2002. relationship between microcystin in drinking water and colorectal cancer. biomed. environ. sci. 15 166-171. no nco mm er cia l u se on ly layout 1 article introduction phytoplankton are microscopic algae and cyanobacteria, but despite being tiny, their intraand interspecific range in size is huge. interspecific linear cell sizes span 5 orders of magnitude, and cell volumes more than 8 orders of magnitude (finkel et al., 2010; ryabov et al., 2021). the amazing morphological phytoplankton diversity in terms of cell size and shape represents a successful adaptive strategy addressed at facing pelagic environment variability (naselliflores, 2021). specifically, cell size is the dominant trait affecting the physiology and ecological function of phytoplankton, including metabolic rate (growth, photosynthesis, respiration), light acquisition, nutrient uptake, sinking rate from the illuminated dominance of small-sized phytoplankton in a mediterranean eutrophic coastal lagoon silvia pulina1,2, cecilia teodora satta1,2*, bachisio mario padedda1,2, jacopo culurgioni3, riccardo diciotti3, nicola fois3, antonella lugliè1,2 1university of sassari, department of architecture, design and urban planning, sassari; 2national biodiversity future center (nbfc), palermo; 3agris, agricultural research agency of sardinia, sassari, italy abstract the predator-prey relationship is generally size-specific in the pelagic food webs. phytoplankton cell size structure can provide information on the successive levels of consumers and therefore on the energy that can flow towards the top consumers. this work focuses on phytoplankton cell size structure in a coastal lagoon (cabras lagoon, italy) considered one of the most important for fishing productivity in the mediterranean. the inter-annual and seasonal dynamics of picophytoplankton (pico, cell size <3 μm) and utermöhl fraction of phytoplankton (ufp, cell size >3 μm) were considered during almost three years in relation to the temporal dynamics of selected environmental variables and zooplankton. small-sized cells with a mean linear cell size <10 µm and a mean cell volume <103 µm3 mainly represented ufp along the entire study period. this size class contributed the most to total phytoplankton biomass (up to 86%) and density (up to 99%) during the first part of the investigation period. a compositional change was detected: smaller species of chlorophyceae, bacillariophyceae, filamentous cyanophyceae, and autotrophic nanoflagellates thrived in the second part of the study, replacing larger mediophyceae that dominated ufp at the beginning. picocyanobacteria rich in phycocyanin were the dominant taxa of pico along the entire investigation period and this size class contributed the most to total phytoplankton biomass (up to 30%) and density (up to 96%) at the end of the study. the observed shift towards different and even smaller ufp and pico in the second part of the study was most probably due to complex interactions between top-down and bottom-up effects. indeed, an increased temperature, a decreased salinity and decreased concentrations of nutrients (mainly ammonium and orthophosphate), as well as an increased grazing pressure of rotifers on the larger mediophyceae were simultaneous with the changes detected in phytoplankton. the obtained results highlight a longer planktonic trophic web in cabras lagoon that includes small phytoplankton at the base, ciliates, rotifers, and copepods. this suggests low energy availability for planktivorous fish, with possible future relevant consequences for fishing activities in this coastal lagoon. corresponding author: cecilia teodora satta, university of sassari, department of architecture, design and urban planning, via piandanna 4, 07100 sassari, italy; nbfc, national biodiversity future center, palermo 90133, italy. e-mail: ctsatta@uniss.it key words: cell size; pelagic food-webs; coastal lagoons; phytoplankton; zooplankton; mediterranean. authors’ contributions: all the authors made a substantive intellectual contribution. all the authors have read and approved the final version of the manuscript and agreed to be held accountable for all aspects of the work. conflict of interest: the authors declare no potential conflict of interest. funding: none. availability of data and materials: all data generated or analyzed during this study are included in this published article. acknowledgments: the authors thank dr bastianina manca and dr pasqualina farina for analyses of algal nutrients, and fishermen’s consortium of cabras lagoon for logistic support during sampling activities. received: 22 december 2022. accepted: 14 april 2023. publisher’s note: all claims expressed in this article are solely those of the authors and do not necessarily represent those of their affiliated organizations, or those of the publisher, the editors and the reviewers. any product that may be evaluated in this article or claim that may be made by its manufacturer is not guaranteed or endorsed by the publisher. ©copyright: the author(s), 2023 licensee pagepress, italy advances in oceanography and limnology, 2023; 14:11112 doi: 10.4081/aiol.2023.11112 this work is licensed under a creative commons attribution noncommercial 4.0 international license (cc by-nc 4.0). no nco mm er cia l u se on ly s. pulina et al.20 surface layer, and predatory grazing by zooplankton (litchman et al., 2008). for these reasons, litchman et al. (2010) have suggested considering cell size a master trait helpful in the analysis of phytoplankton responses to environmental conditions and their change. for example, being small is a particularly good strategy to deal with nutrient depletion because of the high surface-to-volume ratio that makes phytoplankton cells less diffusion-limited (litchmann et al., 2007). consequently, smallersized phytoplankton generally dominate phytoplankton in the open ocean where nutrients are depleted and primary production is low, while larger-sized phytoplankton are generally more dominant in nutrient richer coastal waters where primary production is higher (cabré et al., 2016; marañón et al., 2007). as primary producers at the base of aquatic food-webs, phytoplankton size structure can impact on grazers and, consequently, on the functioning of aquatic ecosystems and their relative ecosystem goods and services (como et al., 2012, 2018). however, the functioning of plankton communities is complex, and in addition to trophic interactions, various kinds of less studied biological links, such as mutualistic symbiosis, commensalism, allelopathy and competitions, may be relevant (kéfi et al., 2015; pilosop et al., 2017; russo et al., 2022). a sizebased predation has been reported for plankton food chains and it is generally assumed that larger prey is eaten by larger predators: heterotrophic nanoflagellates (cell size <20 μm) feed on picoplankton (cell size <2 μm), microzooplankton (20–200 μm) feed on heterotrophic nanoflagellates and nanoplanktonic phytoplankton, and mesozooplankton (200–2000 μm) feed on microplankton (peter and sommer, 2012; sommer et al., 2017a). this means that the energy flow in marine pelagic food webs is largely based on size specific predator–prey relationships (boyce et al., 2015): a shift to smaller size phytoplankton at the base of the food web results in a lower quality food for mesozooplankton, making the carbon flow between primary producers and mesozooplankton longer, through heterotrophic nanoflagellates and microzooplankton, with relative losses of energy, nutrients and carbon which can reduce productivity of higher trophic levels as described by berglund et al. (2007). actually, the trophic diversity of plankton is huge, the functioning of marine pelagic food webs is complex and the efficiency of energy transfer is more variable than previously thought (d’alelio et al., 2016a; eddy et al., 2021). for example, there are pelagic tunicates centimeters in length that filter particles, including phytoplankton, four to five orders of magnitude smaller than themselves (sutherland et al., 2010). both planktonic metazoan and protozoan consumers may change their feeding behaviors based on the available resources, modifying the overall functioning of the plankton food web in different seasons (d’alelio et al., 2016b). seven-fold decrease in phytoplankton biomass translated into only a two-fold decrease in potential planktivorous fish biomass in a plankton community in the gulf of naples (d’alelio et al., 2016a). large temporal fluctuations in the dominance of primary producers (i.e. macroalgae vs. phytoplankton) and, thus, in the available food items, resulted in major changes in the diet and stable isotopic signatures of several primary consumers in the hypertrophic yundang lagoon (xiamen, china; zheng et al., 2020). the economic consequences of plankton composition changes (and food web alterations) are especially important in ecosystems that are strongly exploited for fishing, such as mediterranean coastal lagoons. in general, coastal lagoons are transitional shallow ecosystems highly sensitive to natural and anthropogenic environmental changes, such as climate crises, whose impacts occur at global and local scales (barbosa et al., 2010; newton et al., 2014). coastal lagoons function as tightly linked benthic-pelagic systems and, in addition to the environmental stressors and the pelagic biological interactions, phytoplankton structure and dynamics can be strongly influenced by the benthic environment in these ecosystems. benthos can be a strong sink for phytoplankton biomass (for example, grazing by benthic size-specific filter feeders), or a source for meroplanktonic species from germination of resting stages in sediments (cloern and jassby, 2010) and/or for tychoplanktonic species removed from substrates by turbulence (cahoon, 2016). therefore, decoding phytoplankton spatial and seasonal dynamics is challenging in heterogeneous and dynamic ecosystems. in this regard, the analyses of multiannual data turned out to be a useful instrument for disentangling the intrinsic complexity of phytoplankton dynamics, at least at a local level (morabito et al., 2018; zingone et al., 2010). the scarcity of detailed observations, and the high variability within lagoons, seriously limited the establishment of general models of planktonic succession in lagoons as has been done for other marine and freshwater environments (gilabert, 2001). to date, the environmental and biological factors affecting the entire plankton community composition and size structure in coastal lagoons have been little studied, preventing a deep understanding of plankton food web functioning in these ecosystems (trombetta et al., 2021). mediterranean lagoons are considered particularly vulnerable environments that deserve attention as the mediterranean region is deemed as one of the most sensitive areas regarding on-going global warming and increased extreme climate events (ferrarin et al., 2014; ipcc, 2021). temperature rise is expected to favour the selection of smaller-sized phytoplankton with profound consequences for the aquatic food web structure and efficiency from the very basis to the upper trophic levels (boyce and worm, 2015; polovina et al., 2012; sommer et al., 2017b). however, local dynamics are strongly site-specific, and processes within each region can modulate the overall patterns observed at a global level (chust et al., 2014; lomas et al., 2022). a significant temporal variation in the trophic levels of dominant fish strongly related to fish size has been demonstrated in cabras lagoon, the largest coastal lagoon in sardinia (italy, mediterranean sea) and one of the most important for fishing productivity in the mediterranean basin (como et al., 2018). cabras lagoon is known to be a highly complex eutrophic ecosystem (magni et al., 2022). consistently, very high δ15n values in the small size seston (<55 µm), likely related to a high input of anthropogenic nitrogen from a large catchment area, resulted in a 15n enrichment maintained from primary producers up to the predators at higher trophic levels (como et al., 2012). in addition, studies on ecological data collected since 1999 showed the affirmation of very small-sized phytoplankton (linear cell size <10 µm) in cabras lagoon (padedda et al., 2012; pulina et al., 2011, 2012). to verify if it was only an occasional feature of phytoplankton in this ecosystem, the present work focused on the size structure of phytoplankton from cabras lagoon. two different phytoplankton size classes, the picophytoplankton (cell size <3 μm; pico) and the utermöhl no nco mm er cia l u se on ly dominance of small-sized phytoplankton in a mediterranean eutrophic coastal lagoon 21 fraction of phytoplankton (cell size >3 μm; utermöhl fraction of phytoplankton, ufp), were studied over three additional years (2017, 2018, 2019) not analyzed in previous studies. interannual and seasonal phytoplankton dynamics were related to the temporal variation of environmental conditions in the lagoon and, for the first time in this ecosystem, to the temporal dynamics of zooplankton. we aimed: i) to confirm the dominance of small-sized phytoplankton in the lagoon also during the latest investigation period, suggesting that it is an intrinsic aspect characterizing phytoplankton of this ecosystem that reveals a low energy-efficient trophic web; ii) to detect which environmental variables affected significantly the temporal dynamics of the two phytoplankton size classes (pico and ufp); iii) to assess if and how zooplankton grazing activity may have affected the temporal dynamics of the two size classes in the considered years. materials and methods study area cabras lagoon (western mediterranean sea; 39°56′37″ n, 08°28′43″ e; figure s1) is a eutrophic lagoon with a mean depth of 1.5 m and about 23.8 km2 large, where phytoplankton are the main primary producers. freshwater originates from a large catchment of about 430 km2 and mostly arrives in the northern part of cabras lagoon, carried by the river rio mare’e foghe. the communication with the adjacent gulf of oristano is on the southern opposite side of the lagoon and it happens mainly via narrow creeks that flow into a southernmost canal. because of an ecological dataset available since 1999 (padedda et al., 2010, 2012; pulina et al., 2011, 2012, 2016, 2020; satta et al., 2014), cabras lagoon has been included in the “marine ecosystems of sardinia” site of the italian long term ecological research network since 2006 (https ://deims .org/d5071 d21-9c8f-47ccb534-1b116 2a5e0 9c; morabito et al., 2018; pugnetti et al., 2013). samplings and chemical-physical analyses data on phytoplankton and environmental variables derive from monthly samplings carried out over three years, from april 2017 to october 2019. water samples were collected at three sampling stations, located along the natural gradient of salinity: station 1, near the lagoon’s connection with the sea; station 2, in the center; station 3, near the main input of freshwater (figure s1). in situ, water transparency (tra) was measured with a secchi disk, and water temperature (tem) and salinity (sal) were detected using a multiparameter probe (ysi 6600v2). water samples for phytoplankton and nutrients analyses were collected at about 20 cm depth: 250 ml samples for studying ufp (cell size >3 μm) were immediately fixed in 2% acid lugol’s solution, and 100 ml samples for studying pico (cell size <3 μm) and autotrophic nanoflagellates (cell size >3 μm) were immediately fixed with formaldehyde (2%). in the laboratory, orthophosphate (p-po4), ammonium (n-nh4), nitrate (n-no3), nitrite (n-no2), silicate (si-sio4), total phosphorus (tp) and total nitrogen (tn) concentrations were obtained according to the methods of strickland and parsons (1972). zooplankton samples were collected seasonally (in summer and autumn 2017, in winter, spring and summer 2018, in winter 2019) at station 2, with a hand-held conic net (mouth diameter 31 cm, mesh size 10 μm), towed for 30 m at 50 cm depth with speed of 0.8 1 m s-1 and immediately fixed in 70% ethanol. phytoplankton analyses five ml of fixed water samples were analyzed to estimate ufp cell density using utermöhl’s technique (utermöhl, 1958), with an inverted microscope (axiovert 25, zeiss, oberkochen, germany). cell counts were made at magnifications of 200 x and 400 x for more easily identifiable species, and at magnifications of 1000 x for smaller cells from at least 10% of the total bottom area of the settling chamber. non-fixed samples were also observed to facilitate the identification of species. the species were identified according to the taxonomic literature listed in pulina et al. (2012) and grouped at the class level for data analysis. cell sizes of 20 randomly selected individuals of each taxon were measured in each sample. cell volumes were calculated approximating the shape of each taxon to a geometric shape following hillebrand et al. (1999), vadrucci et al. (2013) and “atlas of shapes” powered by lifewatch italy (http://phytobioimaging.unisalento.it/products/atlasofshapes.a spx?id_tipo=0). for each taxon, the cell carbon content was obtained by applying the conversion formulas suggested by menden-deuer and lessard (2000) to the mean cell volume. the biomass of each taxon, in terms of carbon content, was calculated by multiplying cell carbon content by the corresponding cell density. for pico cell counts, from 2 to 5 ml (depending on cell density) of fixed water samples were filtered onto 0.2-μm blackstained polycarbonate membranes (nucleopore). duplicate slides were prepared and observed using a microscope (axiovert 100, zeiss) equipped with green (bp520–560 nm/ft580 nm/lp590 nm) and blue (bp450–490 nm/ft510 nm/lp520 nm) filter sets, at 1000 x magnification. at least 20 random fields of view and a minimum of 100 randomly selected cells of each taxon (picocyanobacteria rich in phycoerythrin, p-pe; picocyanobacteria rich in phycocyanin, p-pc; picoeukaryotes, p-euk) were counted for each slide (macisaac and stockner, 1993). when it was not possible to recognize a specific taxon, cells were counted as picophytoplankton undetermined (picound). total picophytoplankton (pico) counts were the sum of p-pe, p-pc, p-euk, and pico-und. cell sizes of about 200 randomly selected individuals from each pico group were measured on each slide. the cell volume of each group was calculated assuming that the shape of the cell was spherical or cylindrical with hemispheric ends and using the bratbak formulas (1985). to calculate cell carbon content, the conversion factors of 250 fg c μm-3 and of 220 fg c μm-3 were applied to the mean volume of picocyanobacteria and p-euk, respectively (tamigneaux et al., 1995). the biomass of each pico group was obtained by multiplying the cell carbon content by the corresponding cell density. autotrophic nanoflagellates (cell size 3-20 µm) were also counted during pico cell counts. at least 200 randomly selected flagellate cells were counted and measured for each slide. their cell volume and biomass were calculated as described above for ufp. for each sample, total ufp and pico density and biomass no nco mm er cia l u se on ly s. pulina et al.22 were calculated as the sum of the cell density and biomass, respectively, of all taxa and groups observed, including autotrophic nanoflagellates into ufp. since the arithmetic average of individual morphological traits does not properly represent the community structure because common and rare taxa receive the same weight when calculating the average, weighted average of volume was also estimated for the whole ufp community according to kruk et al. (2015). zooplankton analyses in the laboratory, fixed zooplankton samples were fractionated into mesozooplankton (meso) and microzooplankton (micro), using sieves of mesh size 200 μm and 20 μm, respectively, and preserved in fresh 70% ethanol. each subsample was brought to the volume of 10 ml or 5 ml (depending on the richness of the subsample) by removal of the supernatant and analysed under a stereoscope (leica m80, wetzlar, germany) at 20-50 x magnification to perform a qualitative-quantitative analysis by direct enumeration of the organisms in 1 ml aliquots. for each subsample, counted specimens were identified according to available literature (avancini et al., 2006; cicero et al., 2016). the individuals were grouped at different levels: phylum (nematoda, rotifera), subphylum (naupli), class (copepoda, ostracoda, bivalvia, gastropoda, polychaeta, appendicularia), subclass (teleostei), order (decapoda). taxa that never contributed at least 2% to total zooplankton density were not included into the data analyses. metazoan eggs were not considered either, as they are not primary consumers. data analyses monthly data from the three sampling stations of the entire study period were included in all statistical analyses. in addition, seasons were considered as: winter = january – march, spring = april – june, summer = july – september, autumn = october – december. autumn 2019 was only represented by the month october. to ascertain annual, seasonal, and spatial differences in pico density and biomass, in ufp density and biomass, and selected environmental variables (tra, tem, sal, n-nh4, n-no2, n-no3, si-sio4, p-po4), three-way analysis of variance (anova) was performed considering three fixed factors (years, seasons and sampling stations), individually and combined. since autumn 2019 included only the month “october”, one-way anova was also performed to assess significant differences in the same aforementioned biological and environmental variables among october 2017, october 2018, and october 2019, to confirm (or not) what we detected for the season “autumn”. prior to anova analyses, the two assumptions were confirmed for all pooled data: normal distribution (kolmogorov-smirnov test) and homogeneity of variance (bartlett test). all data required logarithmic [ln(x)] transformation to meet anova assumptions. when significant differences in the dependent variables based on factors were observed, post hoc tukey’s pairwise comparisons test was performed. the generalized linear mixed model (glmm) was used (r package mass) to investigate the influence of selected environmental variables related to the total biomass and total density of pico and ufp. this kind of model is an extension of the generalized linear model (glm) and provides a more flexible approach to analyse non-normal data when random effects are present (bolker et al., 2008). two glmm were applied: one model included tem, tra, sal, p-po4, n-no3, nno2, n-nh4 and p-po4 as fixed terms (predictor variables) and the seasonality as the random variable (random term; pinheiro et al., 2012); the other model included the same selected environmental variables and the months as the random term. all the statistical analyses were performed using r 2.15.2 software (r core team, 2020), and an a priori level of significance was established at p<0.05. results pico and ufp inter-annual and seasonal dynamics monthly values of pico and ufp biomass and density along the study period were reported in table s1. considering pico inter-annual dynamics, after an initial peak in spring 2017 and another peak in spring 2018, maximum values of both biomass and density were detected at the end of the study period, from spring to autumn 2019 (figures s2, s3). pico biomass and density observed in 2019 resulted significantly different from the values observed in 2017 and 2018 (figures s2, s3; table 1; tukey’s test: p<0.001). specifically, spring density in 2019 was higher than spring density values in 2018 table 1. results of the three-way anova (f-test and p-value) to assess significant differences in cell biomass and density of picophytoplankton (pico) and utermöhl fraction of phytoplankton (ufp) among years, seasons and sampling stations. interaction between factors (years x seasons, years x stations, years x seasons x stations) are also shown. significant results (p<0.05) are in bold. years seasons stations years x years x seasons x years x seasons stations stations seasons x stations f p f p f p f p f p f p f p cell biomass pico 10.30 <0.001 4.64 0.01 0.01 0.99 3.47 0.01 0.03 0.99 0.11 0.99 0.33 0.97 ufp 2.21 0.12 3.90 0.01 0.18 0.83 3.44 0.01 0.61 0.65 0.64 0.70 0.36 0.96 cell density pico 21.31 <0.001 6.84 <0.001 0.07 0.93 6.66 <0.001 0.15 0.96 0.22 0.97 0.41 0.93 ufp 16.80 <0.001 1.98 0.13 0.02 0.98 2.40 0.04 0.68 0.61 0.42 0.86 0.35 0.96 no nco mm er cia l u se on ly dominance of small-sized phytoplankton in a mediterranean eutrophic coastal lagoon 23 (tukey’s test: p=0.040), and summer and autumn values of both biomass and density observed in 2019 were higher than the values observed in summer and autumn in 2017 and 2018 (tukey’s test: p<0.001). even pico biomass (anova: f=26.21, p=0.002) and density (anova: f=33.48, p=0.001) detected in october 2019 were significantly higher than the biomass (tukey’s test: p=0.025) and density (tukey’s test: p=0.019) observed in october 2018 (tukey’s test: p=0.020) and higher than the biomass (tukey’s test: p<0.001) and density (tukey’s test: p<0.001) observed in october 2017. pico biomass and density varied significantly among seasons of the same year as well (figures s2, s3; table 1). autumn pico biomass and density were higher than winter biomass (tukey’s test: p=0.020) and density (tukey’s test: p<0.001) in 2019, and summer pico density was also higher than winter density in the same year (tukey’s test: p=0.002). spring pico biomass and density were higher than summer biomass (tukey’s test: p=0.030) and density (tukey’s test: p=0.020) in 2017. considering ufp inter-annual dynamics, maximum values of both biomass and density were detected at the beginning of the study period, from summer 2017 to winter 2018 (figures s2, s3). ufp density in summer 2017 was significantly higher than the values observed in summer 2019, and ufp density in autumn 2017 was higher than the values observed in autumn 2018 (figures s2, s3; table 1; tukey’s test: p=0.004 and tukey’s test: p=0.031, respectively). even ufp density in october 2017 was higher than in october 2018 and 2019 with a significant difference only with october 2018 (anova: f=26.39, p=0.002; tukey’s test: p=0.003). ufp biomass detected in winter 2018 was significantly higher than the values of biomass observed in winter 2019 (figures s2, s3; table 1; tukey’s test: p=0.022). in addition, while no significant differences in ufp biomass among autumns of the different years were observed, ufp biomass in october 2017 was significantly higher than the biomass detected in october 2018 and in october 2019 (anova: f=18.39, p=0.005; tukey’s test: p=0.004 and tukey’s test: p=0.03, respectively). ufp biomass and density also varied among seasons of the same year (figures s2, s3; table 1): autumn ufp biomass and density were higher than spring biomass (tukey’s test: p=0.029) and density (tukey’s test: p=0.031) in 2017, and winter ufp biomass was higher than spring biomass in 2018 (tukey’s test: p=0.022). excluding spring 2017, pico contribution to total phytoplankton biomass and density increased throughout the study period, while ufp contribution decreased (figure s4). the highest pico contribution to total phytoplankton biomass (up to 30%) and density (up to 96%) was observed in spring, summer, and autumn 2019. on the contrary, the highest ufp contribution to total phytoplankton biomass (up to 86%) and density (up to 99%) was observed from summer 2017 to winter 2018. significant differences in pico and ufp biomass and density were not detected among sampling stations in cabras lagoon during the investigated years (table 1). pico and ufp composition considering pico composition, p-pe (mean linear cell size 1.16±0.27 µm and mean volume 0.86±0.43 µm3) and p-pc (mean linear cell size 1.23±0.40 µm and mean volume 0.96±0.55 µm3) were observed throughout the investigated period with similar contributions to total biomass and density (figures s5, s6). p-pc dominated in all seasons in 2017 and 2019 with contributions always >60% to pico total biomass and density. p-pe maximum contributions to pico total biomass and density were observed in autumn 2017 and in winter 2018 (>20%). p-euk (mean linear cell size 1.50±0.01 µm and mean volume 1.77±0.01 µm3) were observed only in spring 2018, when dominated pico total biomass and density (85% and 71%, respectively). considering ufp composition (figures s5, s6), trebouxiophyceae (mainly chlorella sp., mean linear cell size 3.84±0.69 µm and mean volume 8.05±0.30 µm3) contributed the most to total ufp biomass in spring 2017 (43%). this class was even more important in terms of density, with a contribution of at least 13% up to a maximum of 41% from spring 2017 to winter 2018. mediophyceae (mainly cyclotella/thalassiosira sp., mean linear cell size 9.74±6.73 µm and mean volume 869.01±1680.44 µm3) dominated the total ufp biomass, especially at the beginning of the study period, from summer 2017 to spring 2018 and in autumn 2018 with a contribution >60%. the contribution of mediophyceae to total phytoplankton biomass and density was lower in the second part of the investigation period, especially from winter 2019 onwards, when a significant presence affirmation of other groups was also detected. autotrophic nanoflagellates (mainly pyramimonadophyceae and cryptophyceae undetermined, mean linear cell size 7.33±0.25 µm and 9.39±3.54 µm, respectively, and mean volume 37.57±8.50 µm3 and 130.68±115.65 µm3, respectively) contributed the most from spring 2018 to the end of the study period, dominating the total ufp biomass in winter (61%) and autumn 2019 (57%). the contribution of bacillariophyceae (mainly pennales undetermined, mean linear cell size 9.01±2.12 µm and mean volume 15.31±2.21 µm3) and of cyanobacteria (mainly pseudanabaena catenata lauterborn, mean linear cell size 4.09±1.00 µm and mean volume 6.82±2.30 µm3) to total phytoplankton biomass was major from summer 2018 onwards (up to 23% in autumn 2019 and up to 18% in summer 2018, respectively), and the contribution of chrysophyceae (mainly kephyrion spirale (lackey) conrad, mean linear cell size 4.33±0.81 µm and mean volume 27.16±12.68 µm3) to total phytoplankton biomass was up to 18% in summer 2019. chlorophyceae (mainly monoraphidium minutum (nägeli) komárková-legnerová, mean linear cell size 3.42±0.66 µm and mean volume 1.82±1.49 µm3) contributed especially in terms of density. contribution was always >20% from autumn 2017 to spring 2019 and this class dominated the total ufp density in spring 2017 (57%), in autumn 2018 (62%), and in autumn 2019 (78%). as described above, the most abundant ufp taxa detected in the cabras lagoon throughout the investigated period were small in linear cell size (<10 µm) and in mean cell volume (<103 µm3; figure s7). considering all ufp taxa observed during the entire study period, those proliferating with cell densities >107 cell l-1 were those with a mean cell volume <1000 µm3, and taxa proliferating with cell densities >108 cell l-1 were those with a mean cell volume <100 µm3 (figure s8). in addition, the observed change in ufp composition resulted in a reduction in the ufp community volume in the second part of the investigation period (figure s9). no nco mm er cia l u se on ly s. pulina et al.24 inter-annual and seasonal environmental conditions monthly values of the selected environmental variables along the study period were reported in table s2. according to the glmm with seasonality as a random term, among the selected environmental variables, temp, sal, n-nh4, and p-po4 significantly influenced the seasonal dynamic of both pico and ufp throughout the multiannual period in cabras lagoon, but exactly in an opposite way: the environmental variables that positively influenced one phytoplankton size class, negatively influenced the other, and vice-versa (table 2). the same findings came out from the glmm with monthly as a random term, but in this case, n-nh4 affected significantly only ufp dynamics (table s3). considering only the environmental variables affecting significantly phytoplankton throughout the multiannual period in cabras lagoon, the values of sal and p-po4 differed significantly among sampling stations (table 3). these differences never occurred in the same season of the same year, according to tukey’s test. looking at the inter-annual and seasonal dynamics, summer and autumn temp values increased during the years, and autumn 2019 was statistically warmer than autumn 2017 (figure s10; table 3; tukey’s test: p=0.002). october 2019 was significantly warmer than october 2017 and october 2018 (anova: f=303.3, p<0.001; tukey’s test: p<0.001). spring and summer temp were significantly higher than autumn in 2017 (tukey’s test: p<0.001; winter 2017 data are not available) and winter in 2018 (tukey’s test: p<0.001). instead, summer and autumn were significantly warmer than winter in 2019 (tukey’s test: p<0.001). sal in 2017 was significantly higher than sal detected in 2018 and 2019 (figure s10; table 3; tukey’s test: p<0.001). specifically, sal observed in spring, summer and autumn 2017 was higher than the values observed in the same seasons of 2018 (tukey’s test: p<0.001). sal in october 2017 was significantly higher than sal in october 2018 and in october 2019 (anova: f=31.5, p=0.001; tukey’s test: p=0.001 and tukey’s test: table 2. results (t-value and p-value) of the generalized linear mixed-effect model (glmm) applied on the picophytoplankton (pico) and utermöhl fraction of phytoplankton (ufp) total cell density and on the pico and ufp total biomass along the study period (2017, 2018, 2019) in cabras lagoon. the selected environmental variables (secchi disk transparency, tra; water temperature, temp; salinity, sal; ammonium, n-nh4; nitrate, n-no3; nitrite, n-no2; orthophosphate, p-po4; reactive silica, si-sio4) were used as fixed terms and the seasonality was used as random term. significant effects (p<0.05) are in bold. cell biomass cell density pico ufp pico ufp t p t p t p t p tra -0.365 0.716 0.911 0.366 1.003 0.320 -0.825 0.412 temp 0.849 0.399 -3.692 0.000 1.820 0.044 -3.971 0.000 sal -2.452 0.017 6.842 0.000 -1.159 0.251 6.677 0.000 n-nh4 2.418 0.019 -2.995 0.004 1.045 0.300 -1.589 0.117 n-no3 0.906 0.368 -1.027 0.308 -0.038 0.970 -0.883 0.381 n-no2 -1.820 0.074 1.942 0.057 -1.161 0.250 1.510 0.136 p-po4 -2.845 0.006 1.425 0.159 -2.852 0.006 2.024 0.047 si-sio4 0.182 0.856 0.841 0.404 1.237 0.221 -0.700 0.487 table 3. results of the three-way anova (f-test and p-value) to assess significant differences in the selected environmental variables (secchi disk transparency, tra; water temperature, temp; salinity, sal; ammonium, n-nh4; nitrite, n-no2; nitrate, n-no3; orthophosphate, p-po4; reactive silica, si-sio4) among years, seasons and sampling stations. interaction between factors (years x seasons, years x stations, seasons x stations, years x seasons x stations) are also shown. significant results (p<0.05) are in bold. years seasons stations years x years x seasons x years x seasons stations stations seasons x stations f p f p f p f p f p f p f p tra 6.03 0.00 9.41 <0.001 0.18 0.83 3.42 0.01 1.73 0.16 1.46 0.22 1.12 0.38 temp 1.16 0.32 54.03 <0.001 0.19 0.83 4.52 0.00 0.05 0.99 0.09 0.99 0.13 0.99 sal 67.83 <0.001 6.09 <0.001 7.14 0.00 10.14 <0.001 0.73 0.58 0.25 0.96 0.36 0.96 n-nh4 23.09 <0.001 3.12 0.03 1.99 0.15 2.32 0.06 0.30 0.87 0.31 0.93 0.23 0.99 n-no2 4.91 0.01 6.04 0.00 1.68 0.20 5.69 <0.001 0.26 0.90 0.87 0.52 0.37 0.95 n-no3 4.51 0.02 16.97 <0.001 6.65 0.00 4.84 0.00 1.16 0.34 6.08 <0.001 1.06 0.41 p-po4 2.64 0.06 0.92 0.40 7.48 0.00 0.15 0.99 0.05 0.99 2.40 0.05 0.05 0.99 si-sio4 12.45 <0.001 19.85 <0.001 10.56 <0.001 0.74 0.62 4.86 0.00 0.11 0.98 0.54 0.85 no nco mm er cia l u se on ly dominance of small-sized phytoplankton in a mediterranean eutrophic coastal lagoon 25 p=0.007, respectively). sal in spring 2017 was also higher than sal in spring 2019 (tukey’s test: p=0.002). sal in winter 2018 was higher than sal in winter 2019 (tukey’s test: p<0.001). considering seasonal intra-annual differences, sal in spring and summer was higher than sal in winter in 2019 (tukey’s test: p<0.001); sal detected in autumn was significantly higher than in winter (tukey’s test: p<0.001) and spring (tukey’s test: p=0.03) in 2019. regarding algal nutrients, n-nh4 values in 2017 were significantly higher than values observed in 2018 (figure s10; table 3; tukey’s test: p<0.001) and 2019 (tukey’s test: p<0.001). specifically, values observed in spring 2017 were significantly higher than values observed in spring 2018 (tukey’s test: p<0.001) and 2019 (tukey’s test: p<0.001); values observed in autumn 2017 were significantly higher than values observed in autumn 2018 (tukey’s test: p=0.01). even the concentrations detected in october 2017 were only significantly higher than the concentrations detected in october 2018 (anova: f=29.02, p=0.002; tukey’s test: p=0.001). n-nh4 values varied significantly among seasons only in 2017, with spring values higher than summer ones (tukey’s test: p=0.04). p-po4 values detected in 2017 were significantly higher than the values observed in 2018 and 2019 (figure s10; table 3; tukey’s test: p<0.001). specifically, p-po4 values observed in summer and spring 2017 were significantly higher than values observed in summer (tukey’s test: p<0.001) and spring 2019 (tukey’s test: p=0.04). also, p-po4 in summer 2018 was significantly higher than p-po4 in summer 2019 (tukey’s test: p<0.001). p-po4 values varied significantly among seasons only in 2018, with winter values lower than summer values (tukey’s test: p<0.001). zooplankton temporal dynamics and composition zooplankton density was <1 ind l-1 until winter 2018 and increased strongly in spring 2018 (maximum peak of 2.59 ind l-1) and in summer 2018 (figure s11). the meso fraction dominated the total zooplankton density from autumn 2017 (60%) to spring 2018 (100%), while the micro fraction dominated in summer 2017 (70%), in summer 2018 (90%), and in winter 2019 (60%) (figure s11). regarding zooplankton composition, polychaeta larvae represented zooplankton in both meso and micro fractions in summer 2017 (100% to the total zooplankton density) and contributed 63% to the total micro density in autumn 2017 as well (figure s12; table s4). copepods (mainly adults and copepodite stages) dominated total meso density from autumn 2017 (91%) to spring 2018 (85%) and contributed up to 50% (mainly as copepodites and naupli) to the total micro density in winter 2018, showing the same contribution of crustacea naupli. rotifera contributed 44% to the total meso density in winter 2018 and dominated both meso and micro in summer 2018 (88% and 98%, respectively) and in winter 2019 (56% and 89%, respectively). nematoda contributed 4% to the total micro density in autumn 2017. discussion in this work, the temporal dynamics of two phytoplankton size classes, pico (cell size <3 µm) and ufp (cell size >3 µm), were investigated throughout a multiannual period of almost three study years, from april 2017 to october 2019, in a eutrophic mediterranean lagoon. results showed that the size structure of phytoplankton changed strongly, alternating periods of prevalence of ufp on pico and vice versa: ufp dominated at the beginning of the study, especially from summer 2017 to winter 2018; pico dominated the total phytoplankton density in spring 2017 and from summer 2018 to the end of the study period. pico maximum cell density peak of 1.87x109 cell l-1 observed in summer 2019 doubled the ufp maximum peak of 0.95x109 cells l-1 observed in autumn 2017. the contribution of pico to total phytoplankton biomass was clearly lower, up to 30% in spring 2017 and in autumn 2019 when maximum peaks of about 300 µg c l-1 were detected. the pico community was mainly composed of picocyanobacteria, with a higher contribution of p-pc for a longer period compared to the contribution of p-pe. this is in line with the known preference of p-pc for salinity <25 (bec et al., 2011; xia et al., 2017), normally measured in cabras lagoon (pulina et al., 2012). such a relevant presence of pico in cabras lagoon agrees with other studies that have indicated increased pico presence in other mediterranean eutrophic coastal lagoons (bec et al., 2011; collos et al., 2009; pulina et al., 2018; sorokin et al., 2004). pico are well known to be the main contributors of biomass and primary productivity in oligotrophic regions, such as oceanic ecosystems (wang et al., 2022), whereas they have been little investigated in coastal lagoons so far. for this reason, our findings add new ecological information on this component, underlining the necessity of considering them even in studies on phytoplankton from eutrophic transitional environments. the pico cell density detected in cabras lagoon was higher than the values detected previously in other eutrophic sardinian lagoons (up to 108 cells l-1; pulina et al., 2017 and 2018), and closer to values detected in other mediterranean coastal lagoons (venice lagoon, italy; sorokin et al., 2004; bagnas lagoon, france; bec et al., 2011), in an australian coastal lagoon (coorong lagoon, shapira et al., 2010) and other brackish waters worldwide (baltic sea; mazur-marzec et al., 2013; neuse river estuary, usa, paerl et al., 2020). comparison of biomass values is harder since pico biomass data are scarce in the literature. pico biomass values observed in cabras lagoon in the current work were much higher than the values detected previously in the other sardinian lagoons (up to 60 µg c l-1; pulina et al., 2017 and 2018) and much higher than the few values available for coastal waters (about 10 µg c l-1 in chesapeake bay, atlantic ocean; cai et al., 2010, and in ionian sea; caroppo et al., 2014) and for brackish waters (up to 180 µg c l-1 in baltic sea; mazurmarzec et al., 2013). in the present study, small-sized cells mainly represented ufp along the entire study period: they were <10 µm in mean linear cell size and <103 µm3 in mean cell volume. an ufp compositional change was also observed from spring 2018 onwards: the increased abundance of pico at the end of the study was accompanied by the increased abundance of smaller species of chlorophyceae (monoraphidium minutum) and bacillariophyceae (pennales undetermined), smaller autotrophic nanoflagellates, and a smaller filamentous species of cyanobacteria (pseudanabaena catenata), which replaced the larger mediophyceae (cyclotella/thalassiosira sp.) which dominated at the beginning of the study when the maximum ufp abundances were observed. already published results on no nco mm er cia l u se on ly s. pulina et al.26 ecological data collected previously reported small-sized phytoplankton in cabras lagoon (padedda et al., 2012; pulina et al., 2011, 2012, 2016). specifically, cyanobacteria of picoplanktonic cell size (about 3 µm) were the most abundant during the ecological studies performed in the periods 20002002 and 2007-2009 in this ecosystem, together with other small species (cell size 3-10 µm) of cyclotella, fragilaria, thalassiosira, chlorella and with monoraphidium minutum among the most abundant taxa. the size and taxonomic structure of phytoplankton detected in previous studies were confirmed by data reported in the present work obtained with a more appropriate method of analysis applied to investigate pico, the smallest fraction of phytoplankton. hillebrand et al. (2022) reported that phytoplankton cell volume around 100 µm3, such as the main representative mean cell volume of the size class ufp in cabras lagoon, is a very interesting ecological trait: carbon fixation was, in fact, reported to be maximum at cell sizes around 100 μm3, leading to the lowest respiration to photosynthesis ratio at this value of cell volume. in the present work, inter-annual and seasonal dynamics of pico and ufp were related to the temporal dynamics of selected environmental variables, and of microand meso-zooplankton for the first time in cabras lagoon. october 2019 (autumn 2019 in this study) was significantly warmer than october 2017 and october 2018 in cabras lagoon and it was characterised by temperatures very close to the summer values. in addition, with the narrowest range of temperature variation, from 24.3 °c to 27.9 °c, the summer 2019 plus october 2019 was the longest warm period in the present study. changes in species composition with a shift towards smaller-sized organisms, as well as decrements in the average cell size of individual species, have been predicted as a response to ocean warming (sommer et al., 2017b). a strong positive correlation between temperature and small phytoplankton, even among temperature and picocyanobacteria, has been extensively reported in the literature in relation to a direct effect of temperature or due to an indirect effect mediated by a lower nutrient availability in the water and/or a more intense zooplankton grazing activity (collos et al., 2009; d’alelio et al., 2020; peter and sommer, 2012, 2013). in an experimental activity, a plankton community from cabras lagoon was incubated in a laboratory excluding mesozooplankton and it was exposed to heating (pulina et al., 2020). a decreased abundance of pico was detected under warming together with a significant change in ufp taxonomic composition, consisting of the increase of the smaller chlorella sp., which replaced the larger cyclotella sp. and thalassiosira sp. it was related to a direct and positive effect of heating on ciliate abundance in the absence of their predators (mesozooplankton), as well as a taxonomic composition shift of ciliates with the affirmation of ciliates scuticociliatida that preyed upon nanoflagellates and pico. indeed, studies on eutrophic areas show grazing as one of the main drivers behind picocyanobacteria seasonal dynamics, surpassing the role of temperature and nutrient availability in structuring communities (caroppo, 2015). interpreting phytoplankton dynamics in relation to environmental changes in field-based study is tricky, because a lot of different variables including both environmental variables (at local and global scale) and biological relationships can interact and overlap with each other in nature. in the present work, the period summerautumn 2019 (maximum pico contribution) was the longest warm period, and values of salinity and of n-nh4 and p-po4 were lower than the values observed in the summer 2017-winter 2018 period (maximum ufp contribution). according to statistics, smaller ufp species and pico benefited from these conditions of warmer water and lower salinity and n-nh4 and p-po4 concentrations at the end of the study period. small-cell sized phytoplankton are expected to be favoured under lower nutrient concentration because of a higher surface area to volume ratio, a smaller diffusion boundary layer that enables rapid nutrient exchange through the cell surface, and a better competitive ability in nutrient utilisation compared to large-cell sized phytoplankton (harris, 1986; litchman and klausmeier, 2008; van de whaal et al., 2020). temperature, salinity and nnh4 concentrations, strictly related to the freshwater flows through the lagoon, were identified as the main drivers of plankton community dynamics investigated in the hypersaline coorong lagoon (south australia): during high freshwater flow, phytoplankton–zooplankton interactions dominated the food web, while at low flow, bacteria, viruses, and nano/picoplankton interactions were more dominant (hemraj et al., 2017). mesozooplankton (mainly copepods) are well known to feed on microzooplankton and phytoplankton >10 µm in linear cell size (especially diatoms large 20–200 µm) and >500 µm3 in cell volume (lewandowska and sommer, 2010; lewandowska et al., 2014; peter and sommer, 2012). pico, nanoflagellates, flagellates >100 µm, and filamentous cyanobacteria are instead well-known to be inedible for copepods (paul et al., 2021). among smaller zooplankton, rotifers are generally under-represented in zooplankton feeding studies, they are more abundant in freshwaters, and they prey upon bacteria, heterotrophic flagellates, small ciliates and phytoplankton (bonecker and aoyagui, 2005; sweeney et al., 2022). rotifers may be eaten by copepods but can also compete with copepods for the same prey (fussman, 1996). instead, ciliate feeding habits were studied in depth: they are recognized to prey mainly upon picoand nanoplanktonic cell size, but they are also able to eat larger phytoplankton, such as many diatoms (johansson et al., 2004; strom et al., 2001). therefore, according to the literature, the cell size and taxa composition structure of phytoplankton observed in the present work in cabras lagoon was not palatable to copepods, and they may have been more likely eaten by microzooplankton. unfortunately, apart from rotifers, naupli and small copepode stages, the available data-set does not provide deeper insights into microzooplankton, and data on ciliates and heterotrophic phytoplankton are not available in the current work. a significant role of ciliates can only be hypothesized in cabras lagoon also according to the results described by pulina et al. (2020) which demonstrated a significant grazing activity of ciliates on phytoplankton in this ecosystem. looking at the temporal dynamic of phytoplankton and zooplankton in the present study, the strong decrement of ufp abundance and the shift in ufp composition towards smaller ufp taxa started in spring 2018, just when the maximum peak of mesozooplankton copepods was detected in the lagoon. copepods may have prayed upon microzooplankton since microzooplankton was not detected at this time in the lagoon. no nco mm er cia l u se on ly dominance of small-sized phytoplankton in a mediterranean eutrophic coastal lagoon 27 in turn, the increased abundance of the smallest phytoplankton may have been determined by the reduced grazing activity of microzoopankton on them. rotifers appeared in cabras lagoon in winter 2018, just when salinity started to decrease, confirming the strong physiological constraints of saltwater on most rotifers, and that they are particularly sensitive to salinity changes, as highlighted by suikkanen et al. (2013). in a study where plankton composition was investigated in three different north african coastal lagoons, rotifers dominated the zooplankton biomass in the less salty lagoon (lake manzala, egypt, mean salinity of 10.85; ramdani et al., 2009). the peaks of maximum abundance of rotifers observed in summer 2018 and in winter 2019 in cabras lagoon coincided with a strong decreased abundance of mediophyceae, suggesting an intense rotifer grazing pressure on this phytoplankton group, while smaller phytoplankton thrived. this suggests a preferred feeding of rotifers on cyclotella/thalassiosira sp., but not on the smaller phytoplankton such as monoraphidium minutum, small pennales undetermined, autotrophic nanoflagellates and the filamentous pseudanabaena catenata. since the abundance of these smaller phytoplankton started to increase from spring 2018 onwards, just when the highest abundances of copepods first and rotifers then were detected, we can hypothesize that both copepods and rotifers preyed upon ciliates as well, reducing their grazing pressure on the smaller phytoplankton. in mar menor lagoon (spain), a mediterranean coastal lagoon where the composition and temporal dynamics of both phytoplankton and zooplankton were investigated, copepods abundance decreased in late summer when a maximum peak of ciliates and of larger phytoplankton were also observed. on the contrary, the abundance of autotrophic nanoflagellates decreased in this period, suggesting a preferred feeding for copepods on large phytoplankton and ciliates, and for ciliates on the autotrophic nanoflagellates (gilabert, 2001). in cabras lagoon, the longer warm period in 2019 may have kept high the grazing rates of copepods and/or rotifers on ciliates whose decreased grazing pressure on pico may explain the strong affirmation of pico from spring 2019 onwards, together with the decrement in salinity and concentrations of n-nh4 and p-po4. indeed, the literature reported extensively a direct positive effect of experimental warming on the planktonic toppredators (mesozooplankton or microzooplankton when mesozooplankton are absent) and indirect cascade effects in the trophic web on phytoplankton (lewandowska and sommer, 2010; lewandowska et al., 2014; sommer et al., 2007; vidussi et al., 2011). conclusions the dominance of phytoplankton large <10 µm in mean linear cell size and <103 µm3 in mean cell volume in cabras lagoon was highlighted in this work. according to our findings, we retain that both environmental variations and changes in zooplankton composition affected synergistically phytoplankton in cabras lagoon during the study period 20172019. specifically, higher temperature and lower salinity and n-nh4 and p-po4 values, together with a higher grazing pressure of rotifers on the larger ufp species and very likely of copepods and rotifers on ciliates determined a shift towards the affirmation of even smaller ufp taxa and of pico. this framework suggests a longer planktonic trophic web in cabras lagoon, including copepods, rotifers, ciliates, and small phytoplankton that may translate into a low availability of energy for planktivorous fish. references avancini m, cicero am, di girolamo i, et al., 2006. guida al riconoscimento del plancton dei mari italiani. volume ii: zooplancton neritico-testo. icram, 198 pp. available from: https://www.isprambiente.gov.it/it/pubblicazioni/manuali-elinee-guida/guida-al-riconoscimento-del-plancton-dei-mari barbosa ab, domingues rb, galvão hm, 2010. environmental forcing of phytoplankton in a mediterranean estuary (guadiana estuary, south-western iberia): a decadal study of anthropogenic and climatic influences. estuaries coast. 33:324–41. bec b, collos y, souchu p, et al., 2011. distribution of picophytoplankton and nanophytoplankton along an anthropogenic eutrophication gradient in french mediterranean coastal lagoons. aquat. microb. ecol. 63:29–45. berglund j, müren u, båmstedt u, andersson a, 2007. efficiency of a phytoplankton-based and a bacterial-based food web in a pelagic marine system. limnol. oceanogr. 52:121–31. bolker bm, brooks me, clark cj, et al., 2008. generalized linear mixed models: a practical guide for ecology and evolution. trends ecol. evol. 24:127–35. bonecker cc, aoyagui asm, 2005. relationships between rotifers, phytoplankton and bacterioplankton in the corumbá reservoir, goiás state, brazil. hydrobiologia 546:415–21. boyce dg, frank kt, leggett wc, 2015. from mice to elephants: overturning the ‘one size fits all’ paradigm in marine plankton food chains. ecol. lett. 18:504–15. boyce dg, worm b, 2015. patterns and ecological implications of historical marine phytoplankton change. mar. ecol. prog. ser. 534:251–72. bratbak g, 1985. bacterial biovolume and biomass estimation. appl. environ. microbiol. 49:1488–1493. cabré a, shields d, marinov i, kostadinov ts, 2016. phenology of size-partitioned phytoplankton carbon-biomass from ocean color remote sensing and cmip5 models. front. mar. sci. 3:39. cahoon l, 2016. tychoplankton. in: m.j. kennish (eds.), encyclopedia of estuaries. encyclopedia of earth sciences series. springer, new york, usa. 778 pp. cai h, wang k, huang s, et al., 2010. distinct patterns of picocyanobacterial communities in winter and summer in the chesapeake bay. appl. environ. microbiol. 76:2955–60. caroppo c, 2015. ecology and biodiversity of picoplanktonic cyanobacteria in coastal and brackish environments. biodivers. conserv. 24:949–971. caroppo c, musco l, stabili l, 2014. planktonic assemblages in a coastal mediterranean area subjected to anthropogenic pressure. j. geogr. nat. disasters 4:121. chust g, allen ji, bopp l, et al., 2014. biomass changes and trophic amplification of plankton in a warmer ocean. glob. chang. biol 20:2124–39. cicero f, marino g, nasta e, et al., 2016. biodiversità, habitat no nco mm er cia l u se on ly s. pulina et al.28 pelagico e specie non indigene – mesozooplancton. arpa sicilia st3 area mare. eu marine strategy framework directive 2008/56/ec, 198 pp. cloern je, jassby dj, 2010. patterns and scales of phytoplankton variability in estuarine–coastal ecosystems. estuaries coast. 33:230–41. collos y, bec b, jauzein c, et al., 2009. oligotrophication and emergence of picocyanobacteria and a toxic dinoflagellate in thau lagoon, southern france. j. sea res. 61:68–75. como s, magni p, van der velde g, et al., 2012. spatial variations in δ13c and δ15n values of primary consumers in a coastal lagoon. estuar. coast. shelf sci. 115:300–8. como s, van der velde g, magni p, 2018. temporal variation in the trophic levels of secondary consumers in a mediterranean coastal lagoon (cabras lagoon, italy). estuaries coast. 41:218–32. d’alelio d, libralato s, wyatt t, ribera d’alcalà m, 2016a. ecological-network models link diversity, structure and function in the plankton food-web. sci. rep. 6:1–13. d’alelio d, montresor m, mazzocchi mg, et al., 2016b. plankton food-webs: to what extent can they be simplified? adv. oceanogr. limnol. 7:67–92. d’alelio d, rampone s, cusano lm, et al., 2020. machine learning identifies a strong association between warming and reduced primary productivity in an oligotrophic ocean gyre. sci. rep. 10:3287. eddy td, bernhardt jr, blanchard jl, et al., 2021. energy flow through marine ecosystems: confronting transfer efficiency. trends ecol. evol. 36:76–86. ferrarin c, bajo m, bellafiore d, et al., 2014. toward homogenization of mediterranean lagoons and their loss of hydrodiversity. geophys. res. lett. 41:5935–41. finkel zv, beardall j, flynn kj, et al., 2010. phytoplankton in a changing world: cell size and elemental stoichiometry. j. plankton res. 32:119–37. fussmann g, 1996. the importance of crustacean zooplankton in structuring rotifer and phytoplankton communities; an enclosure study. j. plankton res. 18:1897–915. gilabert j, 2001. seasonal plankton dynamics in a mediterranean hypersaline coastal lagoon: the mar menor. j. plankton res. 23:207–17. harris gp, 1986. phytoplankton ecology: structure, function and fluctuation. chapman and hall, london: 384 pp. hemraj da; hossain a, ye q, et al., 2017. anthropogenic shift of planktonic food web structure in a coastal lagoon by freshwater flow regulation. sci. rep. 7:1–10. hillebrand h, acevedo-trejos e, moorthi sd, et al., 2022. cell size as driver and sentinel of phytoplankton community structure and functioning. funct. ecol. 36:276–93. hillebrand h, dürselen cd, kirschtel d, et al., 1999. biovolume calculation for pelagic and benthic microalgae. j. phycol. 35:403–24. ipcc 2021. sixth assessment report wgii. cross-chapter paper 4: mediterranean region. johansson m, gorokhova e, larsson u, 2004. annual variability in ciliate community structure potential prey and predators in the open northern baltic sea proper. j. plankton res. 26:67–80. kéfi s, berlow el, wieters ea, et al., 2015. network structure beyond food webs: mapping non-trophic and trophic interactions on chilean rocky shores. ecology 96:291–303. kruk c, martínez a, nogueira l, et al., 2015. morphological traits variability reflects light limitation of phytoplankton production in a highly productive subtropical estuary (río de la plata, south america). mar. biol. 162:331–41. lewandowska am, hillebrand h, lengfellner k, sommer u, 2014. temperature effects on phytoplankton diversity – the zooplankton link. j. sea res. 85:359–64. lewandowska am, sommer u, 2010. climate change and the spring bloom: a mesocosm study on the influence of light and temperature on phytoplankton and mesozooplankton. mar. ecol. prog. ser. 405:101–11. litchman e, klausmeier ca, 2008. trait-based community ecology of phytoplankton. annu rev ecol. evol. syst. 39:615–39. litchman, e, de tezanos pinto p, klausmeier ca, et al., 2010. linking traits to species diversity and community structure in phytoplankton. hydrobiologia 653:15–28. litchman e, klausmeier ca, schofield om, falkowski pg, 2007. the role of functional traits and trade-offs in structuring phytoplankton communities: scaling from cellular to ecosystem level. ecol. lett. 10:1170–81. lomas mw, bates nr, johnson rj, et al., 2022. adaptive carbon export response to warming in the sargasso sea. nat. commun. 13:1211. macisaac ea, stockner jg, 1993. enumeration of phototrophic picoplankton by autofluorescence microscopy, p. 187-197. in: p.f. kemp, b.f. sherr, e.b. sherr and j.j. cole (eds.), handbook of methods in aquatic microbial ecology. lewis publisher. 800 pp. magni p, semprucci f, gravina mf, 2022. joint analysis of macrofaunal and meiofaunal assemblages improves the assessment of lagoonal environmental heterogeneity. estuar. coast. shelf sci. 266:107740. marañón e, cermeño, jodríguez j, et al., 2007. scaling of phytoplankton photosynthesis and cell size in the ocean. limnol. oceanogr. 52:2190–8. mazur-marzec h, sutryk k, kobos j, et al., 2013. occurrence of cyanobacteria and cyanotoxin in the southern baltic proper. filamentous cyanobacteria versus single-celled picocyanobacteria. hydrobiologia. 701:235–52. menden-deuer s, lessard ej, 2000. carbon to volume relationships for dinoflagellates, diatoms, and other protist plankton. limnol. oceanogr. 45:569–79. morabito g, mazzocchi mg, salmaso n, et al., 2018. plankton dynamics across the freshwater transitional and marine research sites of the lter-italy network patterns fluctuations drivers. sci. total. environ. 627:373–87. naselli flores l, zohary t, padisák j, 2021. life in suspension and its impact on phytoplankton morphology: an homage to colin s. reynolds. hydrobiologia 848:7–30. newton a, icely j, cristina s, et al., 2014. an overview of ecological status vulnerability and future perspectives of european large shallow semi-enclosed coastal systems lagoons and transitional waters. estuar. coast. shelf sci. 140:95–122. padedda bm, lugliè a, ceccherelli g, et al., 2010. nutrient-flux evaluation by the loicz biogeochemical model in mediterranean lagoons: the case of cabras lagoon (centralwestern sardinia). chem. ecol. 26:147–62. no nco mm er cia l u se on ly dominance of small-sized phytoplankton in a mediterranean eutrophic coastal lagoon 29 paerl rw, rebecca e, venezia re, et al., 2020. picophytoplankton dynamics in a large temperate estuary and impacts of extreme storm events. sci. rep. 10:22026. paul c, sommer u, matthiessen b, 2021. composition and dominance of edible and inedible phytoplankton predict responses of baltic sea summer communities to elevated temperature and co2. microorganisms 9:2294. peter kh, sommer u, 2012. phytoplankton cell size, interand intraspecific effects of warming and grazing. plos one 7:e49632. peter kh, sommer u, 2013. phytoplankton cell size reduction in response to warming mediated by nutrient limitation. plos one 8:e71528. pilosof s, porter ma, pascual m, kéfi s, 2017. the multilayer nature of ecological networks. nat. ecol. evol. 1:1–9. pinheiro j, bates d, debroy s, sarkar d, 2012. r development core team, nlme: linear and nonlinear mixed effects models. r package version 3:1–105. polovina jj, woodworth pa, 2012. declines in phytoplankton cell size in the subtropical oceans estimated from satellite remotely-sensed temperature and chlorophyll, 1998–2007. deep sea res. part ii: 77–80, 82–88. pugnetti a, acri f, bernardi aubry f, et al., 2013. the italian long-term ecosystem research (lter-italy) network: results, opportunities, and challenges for coastal transitional ecosystems. transit. waters bull. 7:43–63. pulina s, padedda bm, satta ct, et al., 2012. long-term phytoplankton dynamics in a mediterranean eutrophic lagoon (cabras lagoon, italy). plant biosyst. 146:259–72. pulina s, padedda bm, sechi n, lugliè a, 2011. the dominance of cyanobacteria in mediterranean hypereutrophic lagoons: a case study of cabras lagoon (sardinia italy). sci. mar. 75:111–20. pulina s, satta ct, padedda bm, et al., 2017. picophytoplankton seasonal dynamics and interactions with environmental variables in three mediterranean coastal lagoons. estuaries coast. 40:469–78. pulina s, satta ct, padedda bm, et al., 2018. seasonal variations of phytoplankton size structure in relation to environmental variables in three mediterranean shallow coastal lagoons. estuar. coast. shelf sci. 212:95–104. pulina s, suikkanen s, satta ct, et al., 2016. multiannual phytoplankton trends in relation to environmental changes across aquatic domains: a case study from sardinia (mediterranean sea). plant biosyst. 150:660–70. pulina s, suikkanen s, padedda bm, et al., 2020. responses of a mediterranean coastal lagoon plankton community to experimental warming. mar. biol. 167:22. r core team, 2020. r: a language and environment for statistical computing. r foundation for statistical computing, vienna, austria. http://www.r-project.org/index.html. ramdani m, elkhiati n, flower rj, et al., 2009. environmental influences on the qualitative and quantitative composition of phytoplankton and zooplankton in north african coastal lagoons. hydrobiologia 622:113–31 russo l, casella v, marabotti a, et al., 2022. trophic hierarchy in a marine community revealed by network analysis on cooccurrence data. food webs 32:e00246. satta ct, anglès s, garcés e, et al., 2014. dinoflagellate cyst assemblages in surface sediments from three shallow mediterranean lagoons (sardinia, north western mediterranean sea. estuaries coast. 37:646–663. schapira m, buscot mj, pollet t, et al., 2010. distribution of picophytoplankton communities from brackish to hypersaline waters in a south australian coastal lagoon. saline syst. 6:2. sommer u, aberle n, engel a, et al., 2007. an indoor mesocosm system to study the effect of climate change on the late winter and spring succession of baltic sea phytoand zooplankton. oecologia 150:655–67. sommer u, charalampous e, genitsaris s, moustaka-gouni m, 2017a. benefits, costs and taxonomic distribution of marine phytoplankton body size. j. plankton res. 39:494–508. sommer u, peter kh, genitsaris s, moustaka-gouni m, 2017b. do marine phytoplankton follow bergmann’s rule sensu lato? biol. rev. 92:1011–26. sorokin pyu, sorokin yui, boscolo r, giovanardi o, 2004. bloom of picocyanobacteria in the venice lagoon during summer–autumn 2001: ecological sequences. hydrobiologia 523:71–85. strickland jdh, parsons tr, 1972. a practical handbook of seawater analysis. fisheries research board of canada, ottawa. 167 pp. strom sl, brainard ma, holms jl, olson mb, 2001. phytoplankton blooms are strongly impacted by microzooplankton grazing in coastal north pacific waters. mar. biol. 138:355–68. suikkanen s, pulina s, engström-öst j, et al., 2013. climate change and eutrophication induced shifts in northern summer plankton communities. plos one 8:e66475. sutherland kr, madin lp, stocker r, 2010. filtration of submicrometer particles by pelagic tunicates. pnas 107:15129–34. sweeney k, rollwagen�bollens g, hampton se, 2022. grazing impacts of rotifer zooplankton on a cyanobacteria bloom in a shallow temperate lake (vancouver lake, wa, usa). hydrobiologia 849:2683–703. tamigneaux e, vasquez e, mingelbier m, et al., 1995. environmental control of phytoplankton assemblages in nearshore waters, with special emphasis on phototrophic ultraplankton. j. plankton res. 17:1421–47. trombetta t, vidussi f, roques c, et al., 2021. co-occurrence networks reveal the central role of temperature in structuring the plankton community of the thau lagoon. sci. rep. 11:17675. utermöhl h, 1958. zur vervollkommung der quantitativen phytolankton-methodik. mitt d internat vereinig f. limnologie 9:1–39. vadrucci mr, mazziotti c, fiocca a, 2013. cell biovolume and surface area in phytoplankton of mediterranean transitional water ecosystems: methodological aspects. transit. waters bull. 7:100–23. van de waal db, litchman e, 2020. multiple global change stressor effects on phytoplankton nutrient acquisition in a future ocean. phil. trans. r. soc. b. 375:20190706. vidussi f, mostajir b, fouilland e, et al., 2011. effects of experimental warming and increased ultraviolet b radiation on the mediterranean plankton food web. limnol. oceanogr. 56:206–18. wang f, wei y, zhang g, et al., 2022. picophytoplankton in the west pacific ocean: a snapshot. front. microbiol. 13:811227. no nco mm er cia l u se on ly s. pulina et al.30 xia x, guo w, tan s, liu h, 2017. synechococcus assemblages across the salinity gradient in a salt wedge estuary. front. microbiol. 8:1254. zheng x, como s, huang l, magni p, 2020. temporal changes of a food web structure driven by different primary producers in a subtropical eutrophic lagoon. mar. environ. res. 161:105128. zingone a, phlips ej, harrison pj, 2010. multiscale variability of twenty-two coastal phytoplankton time series: a global scale comparison. estuaries coast. 33:224–9. online supplementary material: figure s1. study site and location of sampling stations. figure s2. inter-annual and seasonal dynamics of (a) picophytoplankton (pico) and (b) utermöhl fraction of phytoplankton (ufp) total biomass in cabras lagoon throughout the study period (2017, 2018, 2019). *autumn 2019 was represented only by october. figure s3. inter-annual and seasonal dynamics of (a) picophytoplankton (pico) and (b) utermöhl fraction of phytoplankton (ufp) total density in cabras lagoon throughout the study period (2017, 2018, 2019). *autumn 2019 was represented only by october. figure s4. picophytoplankton (pico) and utermöhl fraction of phytoplankton (ufp) contribution to (a) total phytoplankton biomass and (b) density in cabras lagoon throughout the study period (2017, 2018, 2019). *autumn 2019 was represented only by october. figure s5. contribution of (a) each picophytoplankton (pico) group to total picophytoplankton (pico) biomass (p-pe, picocyanobacteria rich in phycoerythrin; p-pc, picocyanobacteria rich in phycocyanin; pico-euk, picoeukaryotes; pico-und, picophytoplankton undetermined), and contribution of (b) each utermöhl fraction of phytoplankton (ufp) group to total ufp biomass (bac, bacillariophyceae; chl, chlorophyceae; chr, chrysophyceae; cya, cyanophyceae; dino, dinophyceae; fla, autotrophic nanoflagellates; med, mediophyceae; tre, trebouxiophyceae) in cabras lagoon along the study period (2017, 2018, 2019). *autumn 2019 was represented only by october. figure s6. contribution of (a) each picophytoplankton (pico) group to total picophytoplankton (pico) density (p-pe, picocyanobacteria rich in phycoerythrin; p-pc, picocyanobacteria rich in phycocyanin; pico-euk, picoeukaryotes; pico-und, picophytoplankton undetermined), and contribution of (b) each utermöhl fraction of phytoplankton (ufp) group to total ufp density (bac, bacillariophyceae; chl, clorophyceae; chr, chrysophyceae; cya, cyanophyceae; dino, dinophyceae; fla, autotrophic nanoflagellates; med, mediophyceae; tre, trebouxiophyceae) in cabras lagoon along the study period (2017, 2018, 2019). *autumn 2019 was represented only by october. figure s7. inter-annual and seasonal dynamics of mean cell volume of all ufp taxa observed in cabras lagoon during the study period (2017, 2018, 2019). taxa abundance is not included. *autumn 2019 was represented only by october. figure s8. scatterplot of mean cell volume (x axis) versus cell density (y axis) of each utermöhl fraction of phytoplankton taxon recorded in cabras lagoon along the study period (2017, 2018, 2019). figure s9. inter-annual and seasonal dynamics of ufp community volume in cabras lagoon during the study period (2017, 2018, 2019). weightedaverage of volume of the entire ufp community is showed. *autumn 2019 was represented only by october. figure s10. inter-annual and seasonal dynamics of selected environmental variables (a, temperature; b, salinity; c, ammonium, n-nh4; d, orthophosphates, p-po4) in cabras lagoon throughout the study period (2017, 2018, 2019). *autumn 2019 was represented only by october. figure s11. inter-annual and seasonal dynamics of (a) mesozooplankton (meso) and (b) microzooplankton (micro) total density and their contribution to total zooplankton density in cabras lagoon throughout the study period (2017, 2018, 2019). figure s12. contribution of (a) each taxon of mesozooplankton (meso) and of (b) microzooplankton (micro) to total zooplankton density (biv, bivalvia veliger; cop, copepoda; nau, crustacea naupli; pol, polychaeta larvae; rot, rotifera; nem, nematoda) observed in cabras lagoon throughout the study period (2017, 2018, 2019). table s1. monthly mean and standard deviation (sd) of picophytoplankton (pico) and utermöhl fraction of phytoplankton (ufp) cell density and biomass along the study period. table s2. monthly mean and standard deviation (sd) of the selected environmental variables (tra, secchi disk transparency; temp, water temperature; sal, salinity; n-nh4, ammonium; n-no3, nitrate; n-no2, nitrite; p-po4, orthophosphate; si-sio4, reactive silica) along the study period. table s3. results (t-value and p-value) of the generalized linear mixed-effect model (glmm) applied on the picophytoplankton (pico) and utermöhl fraction of phytoplankton (ufp) total cell density and on the pico and ufp total biomass along the study period (2017, 2018, 2019) in cabras lagoon. the selected environmental variables (tra, secchi disk transparency; temp, water temperature; sal, salinity; n-nh4, ammonium; n-no3, nitrate; n-no2, nitrite; p-po4, orthophosphate; si-sio4, reactive silica) were used as fixed terms and months were used as random term. significant effects (p<0.05) are in bold. table s4. total density (ind l-1) and relative abundance (%) of main zooplankton groups detected in cabras lagoon. no nco mm er cia l u se on ly layout 1 introduction over the past centuries accelerated land use change and over-enrichment of nutrients mainly associated with urban, agricultural and industrial activities have promoted eutrophication of freshwater ecosystems (chamoglou et al., 2014). recent research suggests that eutrophication and climate change are two processes that increase rates of primary production, shifting algal community towards bloom-forming and cyanobacterial species (o’neil et al., 2012; papadimitriou et al., 2013; gkelis et al., 2014). cyanobacterial harmful algal blooms (or cyanohabs) represent one of the most conspicuous waterborne microbial hazards to human and agricultural water supplies, fisheries production, and freshwater and marine ecosystems (codd et al., 2005; paerl et al., 2011). this hazard results from the production of cyanotoxins, harmful secondary metabolites, such as microcystins (mcs), saxitoxin (stx), anatoxin-a (atx-a), and cylindrospermopsins (cyns), which can have deleterious effects within reservoirs and in downstream receiving water systems during releases (paerl and otten, 2013) and pose a threat for living organisms including humans (testai et al., 2016). numerous factors such as water temperature, light attenuation, vertical water mixing and turbidity, flushing rates, residence time, nutrient levels and ratios affect phytoplankton assemblage and biomass composition (i.e., n2fixing vs non-fixing cyanobacteria) (dokulil and teubner, 2000; reynolds et al., 2002; o’neil et al., 2012; pearl, 2014). the warm mediterranean climate favors cyanobacteria blooms in eutrophic waters, which may start in spring and last until december or even throughout the year in hypertrophic lakes (cook et al., 2004; gkelis et al., 2014). in greece, extensive cyanohabs dominated by microcystis, dolichospermum (anabaena), cylindrospermopsis, aphanizomenon, planktothrix, and limnothrix species occur at eutrophic freshwaters, producing mcs, stx and cyn (gkelis et al., 2005; 2014; 2015; gkelis advances in oceanography and limnology, 2017; 8(1): 33-51 article doi: 10.4081/aiol.2017.6350 this work is licensed under a creative commons attribution-noncommercial 4.0 international license (cc by-nc 4.0). monitoring a newly re-born patient: water quality and cyanotoxin occurrence in a reconstructed shallow mediterranean lake spyros gkelis,1* manthos panou,1 ioannis chronis,1 sevasti-kiriaki zervou,2 christophoros christophoridis,2 korina manolidi,2 chrysoula ntislidou,1 theodoros m. triantis,2 triantafyllos kaloudis,3 anastasia hiskia,2 ifigenia kagalou,4 maria lazaridou1 1school of biology, aristotle university of thessaloniki, 541 24 thessaloniki, greece; 2institute of nanoscience and nanotechnology, ncsr “demokritos”, athens, greece; 3water quality department, athens water supply and sewerage company, athens, greece; 4management body of ecodevelopment area of karla, mavrovouni, kefalovriso, velestino, stefanovikio 37500, greece *corresponding author: sgkelis@bio.auth.gr abstract lake karla (central greece) is a unique example at european scale of a shallow lake ecosystem that was dried in the 1960s and in 2009 started to be restored. the lake is listed in the network of the greek protected areas as it is considered a vital aquatic ecosystem, in terms of biodiversity. it has, however, already been adversely affected by both agricultural and industrial land uses in the surrounding area, leading to eutrophication and shifting algal community towards bloom-forming toxic cyanobacterial species. after repeated heavyblooms, cyanotoxin occurrence and mass fish kills, the local ecosystem management authority has implemented a water quality monitoring program (july 2013 july 2015) to assess environmental pressures and the response of aquatic biota in the lake. microscopic, immunological, and molecular techniques combined with physico-chemical parameters, complemented by liquid chromatography tandem mass spectrometry (lc-ms/ms), were used to monitor cyanobacteria blooms and the associated cyanotoxin production from three different sites in lake karla and from the adjacent kalamaki reservoir. water quality was also assessed by the structure of benthic invertebrate community on the sediment. cyanobacteria were the main phytoplankton component, representing more than 70% of the total phytoplankton abundance; dominant taxa belonged to cylindrospermopsis raciborskii, limnothrix redekei, anabaenopsis elenkinii, and microcystis spp. euglenophytes (euglena), diatoms (nitzschia), and chlorophytes (scenedesmus) were also important phytoplankton constituents. lc-ms/ms confirmed the co-occurrence of microcystins, cylindrospermopsin, saxitoxin, neo-saxitoxin and anatoxin-a. the occurrence of cyanotoxins in relation to the persistent and dominant cyanobacteria and the impact of cyanobacterial harmful algal blooms on the newly constructed lake along with the land uses and the emergent mitigation measures are discussed. key words: cyanobacteria; microcystins; saxitoxin; anatoxin-a; cylindrospermopsin; nutrient loads. received: 17 october 2016. accepted: 8 march 2017. s. gkelis et al.34 and zaoutsos, 2014). at present, prediction, prevention and successful elimination of cyanobacterial blooms are still difficult (pearl, 2014; cirés and ballot, 2016) despite the extensive studies and present knowledge on cyanobacteria ecology (dokulil and teubner, 2000; o’neil et al., 2012; pearl, 2014). lake karla (central greece) is a unique example at european scale of a shallow lake ecosystem that was dried in 1960s and is currently undergoing its final re-construction phase for establishing a ‘new’ ecosystem (sidiropoulos et al., 2012). lake karla is one of the most important environmental projects in the region, possibly in the whole country, that has been planned to reverse the adverse environmental conditions, caused by the lake drainage (loukas et al., 2007). its restoration was considered of high importance by the european union as it would offer multi-services i.e. social, economic and ecological sustainable development to the region and not just creating a new reservoir. lake karla is listed in the network of the greek protected areas (it is a natura site (gr1420004), a ramsar site, and a special protected area site for birds). it has, however, experienced adverse effects already from the first year of its refilling, as toxinproducing cyanobacterial blooms (oikonomou et al., 2012; gkelis and zaoutsos, 2014) and fish mortalities or considerable amounts of mcs in fish species have been detected (papadimitriou et al., 2013). in this work our goal was to assess the water quality in the mediterranean, shallow, eutrophic lake karla and the adjacent kalamaki reservoir (central greece) in relation to critical environmental parameters using a multi-approach methodology. physico-chemical parameters combined with biotic parameters (phytoplankton and benthic macroinvertebrates), immunological (elisa), analytical (lc-ms/ms) and molecular techniques (pcr) were used to monitor cyanobacteria blooms and the associated cyanotoxin production on a seasonal basis for two years. methods study area the study was conducted in lake karla located in central greece (39°29΄02΄΄ν, 22°51΄41΄΄ε). lake karla is an ancient lake known from the homeric epics as lake boebeïs (iliad ii.715), which was dried up in 1962 for reclaiming agricultural land and fight floods and malaria. its refilling started at 2010 and the suggested plan proposed the creation of a reservoir of about 38 km2. it has a 288-km perimeter and occupies ca. 1/4 of the old lake karla which covered 180 km2. the hydrological basin of lake karla covers 1660 km2 and the maximum water volume is estimated at 184,000,000 m3. elevation ranges from 50 m to more than 2000 m, and the mean elevation of the region is about 230 m. it is a shallow lake with a current (2013-2015) maximum water depth of 1-1.5 m. the hydrological regime of the lake is determined by inputs (rainfall on the lake and tributary inflows) and the outputs (evaporation). at present, the main water source for lake karla is pinios river. according to the revised restoration plan, two main ditches transfer the flood runoff of pinios river to the reservoir, as it is located in the lower part of karla basin simulating the pre-disturbance conditions. also, four collector channels concentrate the surface runoff from the higher elevation zones of the watershed and directly divert it into the reservoir. the surface runoff of the lower elevation areas will be pumped into the reservoir. so, the maximum allowable volume of the reservoir will reach up to 180 hm3, but only the 60 hm3 will be available to fulfill irrigation needs of the surrounding agricultures because of the environmental restraints, as the primary service of the reservoir will be the establishment of a new wetland. the lake also receives the surface runoff from the surrounding area which is mainly agricultural and stockbreeding and the inflows of perennial streams that drain from the surrounding mountainous land (sidiropoulos et al., 2012). the lake has no natural outflow and the constructed channel draining to the adjacent pagasitikos bay is closed for the present. kalamaki is the oldest and the largest reservoir in the basin of lake karla. today’s purpose of kalamaki reservoir is to keep water for irrigation of the surrounding agricultural area. this reservoir is not included in national monitoring program (egy, 2013) and it is not characterised as water body in the national inventory. however, it is connected with lake karla with a complex irrigation and drainage network and therefore directly affects the physico-chemical and biological quality of karla’s basin because the water that circulates in the basin, remains for a long time in kalamaki reservoir. sample collection, preparation and chemical analysis sampling was taking place bimonthly for the warm period of the year (may-september) and seasonally for the cold period (november-may) during a two-year survey (july 2013-july 2015). water samples were collected from the whole water column from three sampling stations in lake karla (kl1, kl2, kl3) and from the surface of one offshore station in kalamaki reservoir (kk) (fig. 1) using a 1m-long niskin-type sampler and a plastic 10l vessel, respectively. after august 2013, the water level fell in lake karla (kl1, kl2, and kl3 stations) and varied between 0.8-1.2 m. in lake kalamaki (kk station) depth varied between 0.4-0.9 m. after the water level decrease in karla no vessel could enter the lake, therefore sampling was performed by walking. while sampling, extra care was taken to avoid sediment resuspension (after reaching the sampling point researchers waited 10 min before collecting the samwater quality and cyanotoxins in the reconstructed lake karla 35 ple). water transparency was measured using a 20-cm secchi disk. temperature, ph, conductivity, turbidity, and dissolved oxygen were measured in situ using the aquaread ap-2000 probe (kent, gb). concentrations of total phosphorus (tp) in lake water and nitrate (no3-n), nitrite (no2n), ammonia (nh4-n), and soluble reactive phosphorus (srp) in filtrates were processed and analyzed in situ spectrophotometrically using aqua nova 60a and merck standards. three sub-samples of 500 ml each (two fixed with lugol’s solution and formaldehyde and one retained fresh) were collected in polyethylene bottles and used for microscopic analysis. sub-samples (50-300 ml) were filtered through whatman gf/c filters. filter papers and filtrates were stored at -20°c for subsequent cyanotoxin analysis and dna extraction. the carlson trophic state index (tsi) was used as an additional estimator of eutrophication based on abiotic characteristics of the freshwaters surveyed; it was calculated based on tp using the simplified equation given by cooke et al., (1986): tsi=14.42*ln[tp]=4.15. benthic macroinvertebrates were sampled from the soft bottom of lake karla in three replicates for each sampling station (kl1, kl2, and kl3) with an ekman-birge grab (225 cm2 sampling area). in kalamaki reservoir, benthic macroinvertebrates were collected with a handnet because the substrate was coarse: a 250 mm × 230 mm, d-shaped pond net (0.9 mm mesh size, iso 7828:1985; en 27828:1994) was used according to the semi-quantitative 3-min kick/sweep method (armitage and hogger, 1994). then, they were sieved with a 200-mm mesh and fixed in 10% v/v neutralised formaldehyde. phytoplankton analysis fresh, lugol and formaldehyde preserved samples were examined using a zeiss axio imager z2 (carl zeiss, jena, germany) microscope and an inverted microscope (olympus ix71). phytoplankton species were identified using taxonomic keys (komárek and anagnostidis, 1999, 2005; komárek, 2013). phytoplankton abundance was determined in lugol samples with the utermöhl method (utermöhl, 1958). mean cell or filament volume was calculated using geometric formulae after measuring the dimensions of at least 30 individuals (cells or filaments) of each species with an axio cam mrc5 digital camera (carl zeiss). the calculated biovolume concentrations were expressed as biomass per liter (mg l–1) by assuming a specific density of 1 g cm–3. species and taxonomical groups comprising fig. 1. map of lake karla and kalamaki reservoir showing the four sampling stations (kk, kl1, kl2, kl3). insert, map of greece (solid square indicates the location of the two waterbodies). s. gkelis et al.36 more than 10% (w/w) of the total phytoplankton biomass were considered to be dominant. phytoplankton species were classified to functional groups according to reynolds et al. (2002). cyanobacterial biomass was used to assess the water quality by using: i) the two most relevant grades (‘tolerable’ and ‘bad’ corresponding to >8 and >16 mg l–1 of phytoplankton biomass, respectively) given in padisák et al. (2006) for the big shallow lakes of central europe; and ii) the good/moderate (g/m) grade (corresponding to ≤1 mg l–1 of phytoplankton biomass) given by the mediterranean geographical intercalibration group (jrc european commission, 2009); cyanobacteria biomass accounted for more than 80% of the whole phytoplankton biomass, thus was considered representative. cyanobacterial biomass was also compared to the world health organization guidance levels 2 and 3 for recreational waters (who, 2003), which correspond to 10 and 200 mg l–1, respectively, after converting cyanobacterial cell concentration (cells ml–1) to biovolume (mg l–1) according to bartram et al. (1999). benthic macroinvertebrates macroinvertebrates were sorted and identified to genus or species level. chironomids and oligochaetes were slide-mounted prior to determination. for the identification of chironomids (larvae and pupae) and oligochaetes the keys of wiederholm (1983 and 1986) and of timm (2009) were used, respectively. immature stages were identified by the particular characters of the setae. dna extraction and molecular analyses in order to identify potentially toxic cyanobacteria, different primer pairs, previously described in the literature, were used to detect different gene targets known to be involved either in the biosynthesis of mc, cyn or stx. dna was extracted using the protocol described in atashpaz et al. (2010) for gram negative bacteria, after slicing the two offshore sampling stations (station 3 and station 4) filters with a sterile scalpel. pcr was carried out on the dna extracts using the primer pairs shown in tab. 1 and pcr conditions described in detail by gkelis and zaoutsos (2014). all primer pairs were specific for the gene region each one amplifies. thermal cycling was carried out using an eppendorf mastercycler pro (eppendorf). pcr products were separated by 1.5% (w/v) agarose gel in 1x tae buffer. the gels were stained with ethidium bromide and photographed under uv transillumination. dna extracted from microcystis aeruginosa m6 strain was used as positive control for the amplification of mcya, mcyb and mcye/ndaf gene targets; dna from cylindrospermopsis raciborskii aqs strain was used as positive control for the amplification of the ps (peptide syntethase) and pks (polyketide synthase) genetic determinants; dna from aphanizomenon gracile a040 strain was used as positive control for the detection of sxti target gene (see vasconcelos et al., 2010). dna of control strains was extracted as described earlier from lyophilized biomass provided by prof. vitor vasconcelos (ciimar, university of porto, portugal). cyanotoxin analysis the abraxis microcystin (520011), saxitoxin (52255b), and cylindrospermopsin (522011) microtiter plate kits were used to determine the presence of mcs, stxs, and cyns, respectively in sampling stations kl1, kl3, and kk. toxins from filters (i.e., representing the intracellular fraction) were extracted by eight ml of water in glass tubes, immersed in ice and sonicated for 10 min, after slicing the filters with a sterile scalpel. after sonication, the mixture tab. 1. pcr primers used in the analyses of water samples collected from lake karla and kalamaki reservoir for the detection of genes involved in mc, cyn and stx production. primer target-gene sequence (5’–3’) size (bp) reference mcya-cd1f mcya aaaattaaaagccgtatcaaa mcya-cd1r aaaagtgttttattagcggctcat 297 hisbergues et al., 2003 mcyb2959f tgggaagatgttcttcaggtatccaa mcyb3278r mcyb agagtggaaacaatatgataagctac 320 nonneman and zimba, 2002 pkef1 mcye cgcaaacccgatttacag pker1 cccctaccatcttcatcttc 755 ouahid et al., 2005 hepf tttggggttaacttttttgggcatagtc hepr mcye/ndaf aattcttgaggctgtaaatcgggttt 472 jungblut and neilan, 2006 psm13 ps ggcaaattgtgatagccacgagc psm14 gatggaacatcgctcactggtg 597 schembri et al., 2001 pksk18 cctcgcacatagccatttgc pksm4 pks gaagctctggaatccggtaa 422 fergusson and saint, 2003 sxt1-f sxti gcttactaccacgatagtgctgccg sxt1-r ggttcgccgcggacattaaa 1669 kellmann et al., 2008 water quality and cyanotoxins in the reconstructed lake karla 37 was stirred for 30 min at room temperature, centrifuged for 10 min at 13,000 g and the supernatant was collected. the pellet was resuspended in eight ml of water and re-extracted. the resulting solutions were pooled together, dried in an air stream, and the residue was dissolved in 1 ml milli-q water; then, they were applied to the above mentioned elisa kits following the manufacturer’s instructions. extracellular cyanotoxins were measured by applying the filtrates to elisa. all microtiter plates were read at 450 nm and b/b0 values (%) were calculated. samples with a coefficient of variation percentage of >15% were not accepted. dilutions of extracts until they fit each kit’s standard curve range were preformed when necessary. results are given in μg of toxin-equivalent (eq.) per l of water (e.g., eq. μg mc l–1). the detection limits for mc, stx, and cyn assays, are 0.10 μg l–1, 0.015 μg l–1, and 0.040 μg l–1, respectively. in order to confirm the elisa results and verify the identities of cyanotoxins, 13 out of the previously mentioned filter-extracts (intracellular toxins) were analyzed without further pretreatment, using liquid chromatography-tandem mass spectrometry (lc-ms/ms). determination was carried out on a finnigan tsq quantum discovery max triple-stage quadrupole mass spectrometer (thermo fischer scientific, waltham, ma, usa), equipped with electrospray ionization (esi) source. separation of target analytes was achieved with a finnigan surveyor lc system, equipped with a finnigan surveyor as autosampler (thermo fischer scientific). detection was performed in multiple reaction monitoring (mrm) mode. xcalibur software 2.1 sp 1160 was used to control the mass spectrometer and for data acquisition. the determination of cyn, atx-a, and mcs ([dasp3]mc-rr, mc-rr, mc-yr, mc-htyr, [dasp3]mc-lr, mc-lr, mc-hilr, mc-wr, mc-la, mc-ly, mc-lw, mc-lf) was carried out according to the lc-ms/ms method described by zervou et al. (2017). the same extracts were also analysed for stx and neostx with an in-house developed method (moustakagouni et. al., 2016), using a sequant zic-hydrophilic interaction chromatographιc (hilic) column (150 x 2.1 mm, 3.5 µm column) supplied by merck. the method limits of detection (lod) and method limits of quantification (loq) for each cyanotoxin analyzed referring to 50 ml of water concentrated to a final volume of 1 ml are given in supplementary tab. 1. a representative chromatogram of sample kκ (26-7-2013) is given in supplementary fig. 1. cyanotoxin concentrations are given in μg l–1. statistical analyses a one-way anova test was used to compare the means of physical, chemical and biological parameters among sampling stations, after check for normality (shapiro-wilk test) and equality of variances (levene statistic). spearman’s rank correlation coefficient (r) was used to determine correlation between variables. analyses were performed with spss 23.0 (ibm spss statistics). results physical-chemical parameters there was a strong seasonal cycle of water temperature ranging from 14°c to 31°c except in february 2015 where temperature fell to 5°c (fig. 2a). the ph values ranged from 7.1 to 9.5, with a maximum of 10.6 in september 2014 at kl1 station (fig. 2b). dissolved oxygen varied noticeably in time and space, decreasing to <7 mg l–1 at some stations during the warm months (fig. 2c). conductivity was very high throughout the whole study period, significantly higher (anova, p<0.05) in lake karla compared to kalamaki reservoir (fig. 2d). water transparency was low (<35cm) throughout the whole study period (fig. 2e). turbidity was relatively low until mid-2014 where a sharp peak was recorded, after which it stayed high for the rest of the study period (fig. 2f). with regard to nutrient concentrations, nitrite nitrogen (no2-n) was very low and reached undetectable concentrations in all stations until may 2014, after which it rose and reached a peak of 2.5 mg l–1 at may 2015 (fig. 3a). nitrate (no3-n) and ammonia (nh4-n) nitrogen were very high and exhibited temporal and spatial variation (fig. 3 b,c). ammonia nitrogen was the most important form in the din pool. din concentration ranged from 500 μg l–1 at the beginning of the study to >15 mg l–1 (july-november 2015) (fig. 2d). srp and tp concentrations were high and varied between undetectable concentrations (novembermay 2014) and 3 mg l–1 (in september 2013) (fig. 3 e,f). the din/srp was generally low and below the 10 threshold, except between july-november 2014 (fig. 4a). phytoplankton a total of 44 phytoplankton species were identified in the lake water samples during the study period (tab. 2). cyanobacteria were the taxonomic group with the highest number of species (23), followed by chlorophytes (13), diatoms (6), cryptophytes (1), and euglenophytes (1). cyanobacteria dominated the phytoplankton’s biomass (fig. 5). phytoplankton blooms were observed in lake karla and kalamaki reservoir almost throughout the whole monitoring period with cyanobacteria consisting on average ca. 85% of the total phytoplankton biomass (fig. 5). total cyanobacterial biomass ranged from 15 to 230 mg l–1 reaching its peak at september 2013 (kl1 and kk stations), february 2013 (kk) and july 2014 (kl1) (fig. 6). the biomass temporal variation was not the same in the two years of monitoring: after the warm period of 2013 showed a decrease in november to reach high values durs. gkelis et al.38 fig. 2. physical and chemical parameters in lake karla (kl1, kl2, kl3) and kalamaki reservoir (kk) from july 2013 to july 2015. a) temperature; b) ph; c) dissolved oxygen; d) conductivity; e) transparency; f), turbidity. water quality and cyanotoxins in the reconstructed lake karla 39 ing the cold period (february-may) of 2014. the rapid decrease of temperature in november 2014 combined with the high turbidity coincided with the lower biomass values (except kl1 station) in the warm period in 2014; biomass remained low until the end of the monitoring program in july 2015. biomass was negatively but not highly correlated with ph, do, no3-n, no2-n, and conductivity (supplementary tab. 2). the biomass was significantly higher (p<0.05) in kalamaki reservoir (kk) and lake karla kl1 station compared to kl2 and kl3 stations. all biomass values exceeded the good/moderate and bad threshold for ecological status, as well as the who guidance level 2; during the warm period the biomass also exceeded the who guidance level 3 (fig. 4b). in both waterbodies nitrogen-fixing nostocales cyanobacteria were dominant throughout the study period tab. 2. phytoplankton taxa identified in lake karla and kalamaki reservoir during the study period. taxa cyanobacteria (23) anabaenopsis elenkinii v.v.miller 1923 aphanizomenon sp. aphanocapsa sp. arthrospira sp. chroococcus sp. cuspidothrix issatchenkoi (usacev) rajaniemi, komárek, willame, hrouzek, kastovská, hoffmann et sivonen 2005 cylindrospermopsis raciborskii (woloszynska) seenayya et subba raju 1972 dolichospermum spp. gloeocapsa sp. gomphosphaeria sp. limnothrix redekei (van goor) meffert 1988 merismopedia sp. microcystis aeruginosa (kützing) kützing 1846 microcystis flos-aquae (wittrock) kirchner 1898 microcystis sp. planktolyngbya cf. limnetica planktothrix agardhii (gomont) anagnostidis et komárek 1988 pseudanabaena limnetica (lemmermann) komárek 1974 pseudanabaena mucicola (naumann et huber-pestalozzi) schwabe 1964 radiocystis geminata skuja 1948 snowella litoralis (häyrén) komárek et hindák 1988 sphaerospermopsis aphanizomenoides (forti) zapomelová, jezberová, hrouzek, hisem, reháková et komárková 2012 synechococcus sp. chlorophyta (13) chlamydomonas sp. chlorella sp. chlorococcum citriforme archibald et bold 1970 coelastrum sp. desmodesmus communis (e. hegewald) e.hegewald 2000 dictyosphaerium pulchellum h.c. wood 1873 gloeocystis sp. monoraphidium sp. pediastrum boryanum (turpin) meneghini 1840 selenastrum sp. spirogyra sp. tetraëdron minimum (a.braun) hansgirg 1888 tetrastrum komarekii hindák 1977 diatoms (6) aulacoseira granulatα (ehrenberg) simonsen 1979 cyclotella sp. nitzschia acicularis (kützing) w. smith 1853 nitzschia closterium (ehrenberg) w. smith 1853 stephanodiscus sp. synedra sp. cryptophyta (1) rhodomonas sp. euglenophyta (1) euglena acus s. gkelis et al.40 fig. 3. nutrient loads in lake karla (kl1, kl2, kl3) and kalamaki reservoir (kk) from july 2013 to july 2015. a) nitrite nitrogen (no2-n); b) nitrate nitrogen (no3-n); c) ammonia nitrogen (nh4-n); d) dissolved inorganic nitrogen (din); e) soluble reactive phosphorus (srp); f) total phosphorus (tp). water quality and cyanotoxins in the reconstructed lake karla 41 (figs. 5 and 6). in lake karla, the dominant cyanobacteria were anabaenopsis elenkinii, sphaerospermopsis aphanizomenoides (functional group h1), limnothrix redekei and planktothrix cf. agardhii (functional group s1), and cylindrospermopsis raciborskii (functional group sn). during the warm period microcystis aeruginosa (functional group m) was also dominant in kl2 and kl3, whereas in july 2015 in kl2 there was an almost 100% dominance of dolichospermum (anabaena) cf. smithii (fig. 6). in kalamaki reservoir a. elenkinii, c. raciborskii, and sph. aphanizomenoides were dominant from july 2013 to september 2014; in may 2015 d. cf. smithii became dominant whereas in july 2015 lyngbya sp. occurred and became dominant for the first time (fig. 6). cyanobacteria dominated kalamaki reservoir even in the cold period. a. elenkinii, c. raciborskii, sph. aphanizomenoides, and dolichospermum spp. accounted for the higher biomass in kl1 and kk. apart from cyanobacteria, the chlorophytes fig. 4. dissolved inorganic nitrogen (din) to soluble reactive phosphorus (srp) ratio (a) and cyanobacteria biomass (b) in water samples collected in lake karla (kl1, kl2, kl3) and kalamaki reservoir (kk) from july 2013 to july 2015. dotted lines represent the thresholds for ecological status based on cyanobacterial biomass given by padisák et al. (2006) and the mediterranean geographical intercalibration group (jrc european commission, 2009); long-dashed lines represent who guidance levels for cyanobacterial concentrations in recreational waters (who, 2003). s. gkelis et al.42 monoraphidium sp. (functional group x1) and tetrastrum komarekii and the diatoms nitzschia acicularis and nitzschia closterium (functional group x1) were occasionally dominant in both water bodies. benthic invertebrates in lake karla, only four taxa (annelida, lymnaea sp., chironominae, and tanypodinae) were recorded; in kalamaki reservoir 3 of the four taxa were found (lymnaea sp. was absent). molecular detection pcr products indicating the presence of mcya, mcyb, and mcye/ndaf genes were obtained for most of the samples tested (tab. 3). only two of the assayed environmental samples gave positive pcr results for each of the psm13/psm14 and pksm4/pksk18 primer pairs, thus, suggesting the presence of cyn producing cyanobacteria. eight water samples, mainly from kalamaki reservoir, dominated by a. elenkinii, sph. aphanizomenoides, and c. raciborskii, gave a pcr product of about 1500 bp using the sxt1f/sxt1r primer pair for the presence of the sxti gene (tab. 3). cyanotoxins mcs were detected by elisa in water samples collected in the warm period, whereas in the cold period they were below detection limit (fig. 7 a,b). intracellular mclr eq. concentration ranged from 1 μg l–1 to 8 μg l–1 (lake karla) or 11.5 μg l–1 (kalamaki reservoir) (fig. 7a) while extracellular mc concentrations exhibited similar seasonal and temporal pattern with concentrations ranging from 1 μg l–1 eq. to 4 μg l–1 (fig. 7b). a maximum total (sum of intracellular and extracellular) mc concentration of 14.4 μg l–1 eq. was recorded at kk station on september 2013. the mc and cyn intracellular concentrations were positively correlated with temperature and srp (supplementary tab. 2). cyn was found in the warm periods at stations kl1 and kk and in november 2013 at kk (fig. 7c). all those samples were dominated by a. elenkinii, sph. aphanizomenoides, and c. raciborskii. stx was found in some samples in the warm period in very low concentrations, close to the detection limit, with the exception of july 2014 in kalamaki reservoir where it reached 0.17 μg l–1 eq. (fig. 7d). in this sample, a. elenkinii, sph. aphanizomenoides, and c. raciborskii were also the dominant species. fig. 5. relative biomass (%) of cyanobacteria and other algae in water samples collected from lake karla and kalamaki reservoir between july 2013 and july 2015. values represent means of the four sampling stations. water quality and cyanotoxins in the reconstructed lake karla 43 lc-ms/ms analysis confirmed the presence of cyn in some cases, with concentrations ranging from 0.5 μg l–1 to 12.1 μg l–1 (tab. 4). it also revealed the presence of atx-a in 3 out of 13 samples analyzed, with concentrations ranging from 0.8 μg l–1 to 5.4 μg l–1 (tab. 4). again, in those samples a. elenkinii, sph. aphanizomenoides, and c. raciborskii were the dominant species. stx was found in all of the samples analysed, whereas neo-stx was found in 7/13 samples. [d-asp3] mc-rr, mc-rr, mc-yr, and mc-lr were found in trace amounts during the warm periods at stations kl1 and kk; in november 2013 at kk, mc-yr and mclr were found in concentrations below the quantification limit (tab. 4). in july 2013 cyn, atx-a, stx and/or neostx, and mcs co-existed in kl1 and kk. ecological classification according to the water frame directive terminology, the ecological status of modified lakes is expressed as ecological potential and the goal is to achieve at least tab. 3. pcr amplification of regions targeting genes responsible for mc (mcya, mcyb, and mcye/ndaf), cyn (ps, pks) and stx (stxi) production, in the water samples collected from lake karla (kl1, kl3) and kalamaki reservoir (kk). collection sampling cyanotoxin genes date station mcya mcyb mcye/ndaf ps pks sxti 26-07-13 kl1 + + + kl2 + + kl3 + + + kl4 + + + 11-09-13 kl1 + + + + kl2 + + + kl3 + + + kk + + + + 22-11-13 kl1 + + + kl2 + + + kl3 + kk + + + 21-02-14 kl1 + + + kl2 + + + + kl3 + + + kk + + + 15-05-14 kl1 + + + kl2 + + + kl3 + + + kk + + + 04-07-14 kl1 + + + kl2 + + kl3 + + + kk + + + 04-09-14 kl1 + + + kl2 + + + kl3 + + + kk + + + 07-11-14 kl1 + + + kl2 + kl3 + + + kk + + + + 21-02-15 kl1 + kl2 kl3 + kk + 15-05-15 kl1 kl2 + + + kl3 kk + + + + + 01-07-15 kl1 + + kl2 + + kl3 + kk + + + + s. gkelis et al.44 good ecological potential. considering the parameters: i) very high phytoplankton biomass (fig. 4b); ii) dominance of cyanobacteria in phytoplankton (fig. 5); iii) functional groups dominating the phytoplankton (fig. 6); iv) low diversity of macroinvertebrate taxa; v) frequency and intensity of water blooms; and vi) presence of multiple cyanotoxins (fig. 7, tab. 4), the ecological potential in lake karla and kalamaki reservoir is classified as poor. discussion this study presents the simultaneous investigation of the phytoplankton, macroinvertebrate community, and the presence of cyanotoxins in relation with key limnological features (nutrients, temperature, ph, etc.) in a recently restored, highly eutrophic to hypertrophic (chamoglou et al., 2014; theologou et al., 2016) shallow mediterranean lake. in lake karla and kalamaki reservoir dense blooms dominated by cyanobacteria were observed throughout the year, in accordance with previous findings in other eutrophic freshwaters of greece (cook et al., 2004; gkelis et al., 2014). mediterranean lakes are subjected to large variations in water level, determined by naturally intraand inter-annual variations in rainfall and groundwater discharge or recharge in alternating drought and wet periods (beklioglu et al., 2007). temperature variations may have considerable further effects on the lake’s ecosystem structure and dynamics; for example, their response to eutrophication seems to be quite different from that of the cold temperate in northern europe: nutrients can limit phytoplankton biomass throughout the year at low latitudes (moustakagouni et al., 2014). prolonged hydraulic retention time because of drought, results in increased salinity with secondary effects on biota, ion toxicity and osmotic stress (jeppesen et al., 2007). such conditions can significantly reduce the resilience of the lake ecosystems affecting also their goods and services (kagalou, 2010). our results indicate that n and p inflows combined with extremely high water retention time in lake karla and kalamaki reservoir (chamoglou et al., 2014) were the main drivers of the phytoplankton succession and dominance. the nostocales dominance can be linked to the nutrient availability: the din/srp ratio was <10 for the most part of the monitoring period, being close to zero in the beginning of the monitoring, which probably triggered the increase of nitrogen-fixing cyanobacteria that dominate under such conditions (villena and romo, 2003; kagalou et al., 2008; kolzau et al., 2014). the cyanobacteria that formed blooms belong to the functional groups h1, s1, and sn (reynolds et al., 2002; mantzouki et al., 2016). l. redekei, p. agardhii, and c. raciborskii (s1 and sn species) are particularly tolerant in low irradiance (mischke and nixdorf, 2003) and this could account for their dominance in the turbid environment and during the cold period. c. raciborskii, particularly, although generally considered to thrive in warm (>25°c) waters (see dokulil, 2016 and references therein), has been reported to grow at 17°c in german waters (mischke, 2003) and recently dokulil (2016) showed its ability to survive in the vegetative form at water temperatures below 12°c. its presence in lake kalamaki in temperatures as low as 12°c may be further evidence of the wide tolerance spectrum of c. raciborskii, which enables it to survive and thrive in novel environments (rzymski and poniedziałek, tab. 4. cyanotoxin occurrence and concentration (intracellular) in water samples collected from lake karla (kl1, kl3) and kalamaki reservoir (kk) as determined by lc-ms/ms analysis. collection sampling toxin concentration (µg l–1) date station cyn atx-a stx neo-stx [d-asp3] mc-rr mc-yr mc-lr mc-rr 26-07-13 kl1 2.6 0.8 <loq <loq n.d. <loq n.d. <loq kk 10.1 5.4 <loq n.d. n.d. <loq n.d. <loq kl3 n.d. n.d. <loq <loq <loq n.d. <loq <loq 11-09-13 kl1 n.d. n.d. <loq n.d. n.d. n.d. <loq n.d. kl3 n.d. n.d. <loq <loq n.d. n.d. n.d. n.d. kk 0.5 n.d. <loq <loq n.d. n.d. n.d. <loq 22-11-13 kl1 n.d. n.d. <loq <loq n.d. n.d. n.d. n.d. kk 1.0 n.d. <loq n.d. n.d. n.d. <loq <loq 21-02-14 kl3 n.d. n.d. <loq <loq n.d. n.d. n.d. n.d. 04-07-14 kl1 12.1 2.2 <loq n.d. n.d. <loq n.d. <loq kk <loq n.d. <loq <loq n.d. n.d. n.d. n.d. 04-09-14 kk <loq n.d. <loq n.d. n.d. n.d. n.d. n.d. 07-11-14 kl1 n.d. n.d. <loq n.d. n.d. n.d. n.d. n.d. loq, limit of quantification; n.d., not detected. water quality and cyanotoxins in the reconstructed lake karla 45 2014). furthermore, s1 and sn species are only sensitive to flushing (mantzouki et al., 2016), which never occurred during the monitoring period or earlier (chamoglou et al., 2014). in shallow mediterranean lakes, cyanobacterial blooms and their long persistence are more expected at low flushing rates (romo et al., 2004). a. elenkinii and sph. aphanizomenoides (h1 species) are only sensitive to low light, low phosphorus and mixing (reynolds et al., 2002; mantzouki et al., 2016), conditions rarely met in lake karla and kalamaki reservoir, enabling them to dominate phytoplankton. l. redekei, p. agardhii a. elenkinii and sph. aphanizomenoides were found dominant also during previous short-term studies (papadimitriou et al., 2013; gkelis and zaoutsos, 2014). microcystis species, usually forming blooms in freshwater of greece (gkelis et al., 2005, 2014, 2015), also present in lake karla’s phytoplankton, never dominated a bloom during the monitoring period, although such blooms have been reported previously (gkelis and zaoutsos, 2014). the high salinity in lake karla could also have affected microcystis and/or other species as it is known that has a negative impact on the diversity of different biota (padisák et al., 2006). since microcystis species are favoured by high irradiance, high temperatures, and they are tolerant to dark anoxic conditions (šejnohová and maršálek, 2012), being sensitive only to flushing (reynolds et al., 2002) or mixing (visser et al., 2016), it possible that microcystis-dominated blooms will occur in the future in lake karla. apart from phytoplankton, only four and three benthic macroinvertebrate taxa were recorded in lake karla and in kalamaki reservoir, respectively, compared to the 12 taxa which had been recorded in the-sole macroinvertebrates’ study existingby ananiadis (1956), showing a serious deterioration of the water quality. a comparison of the benthic fig. 6. biomass of each phytoplankton taxon per sampling period in water samples collected from lake karla and kalamaki reservoir between july 2013 and july 2015. only taxa contributing >2% to the total biomass are shown. s. gkelis et al.46 macroinvertebrate diversity with lake paralimni, another shallow lake that dried up several times because of the drought (lakes network, 2016), using the shannonwiener index (shw) showed that lake paralimni (shw = 1.59) has higher biodiversity than lake karla (shw = 0.73) and kalamaki reservoir (shw = 1.25) (ntislidou, unpuplished data). concerning the cyanotoxin production, the main finding of our study was the occurrence of multiple cyanotoxins in several samples of lake karla and kalamaki reservoir. it is known that multiple types of cyanotoxins can be produced by individual cyanobacterial strains and can co-occur in environmental populations, e.g. mcs plus atx-a, mcs plus atx-a(s), and nodularin plus mcs (see codd et al., 2005). co-occurrence of mcs and cyns has been found in france (brient et al., 2009) whereas co-occurrence of mcs and stxs in brazil (costa et al., 2006). in lake karla, however, we found that mcs, stx, neostx, and atx-a can co-occur, and to the best of our knowledge, this has been reported only once in freshwaters from the midwestern united states (graham et al., 2010). the cyn concentrations we found here are within the fig. 7. cyanotoxin concentrations detected by elisa in water samples collected from lake karla and kalamaki reservoir between july 2013 and july 2015. a) intracellular mcs; b) extracellular mcs; c) cyn; d) stx. water quality and cyanotoxins in the reconstructed lake karla 47 range obtained in previous studies (brient et al., 2009; kokociński et al., 2009; berry and lind, 2010), although a value similar to the maximum concentration detected in lake karla has only been reported in germany (rücker et al., 2007) as dissolved cyn. studies suggest that cyn is dominantly extracellular (bormans et al., 2014) therefore the intracelullar concentrations found here may be an underestimation of the total cyn present in the lakes. stxs, generally reported to occur more rarely in freshwaters, have been found in several countries (testai et al., 2016). stx concentrations in lake karla are similar to those reported by clemente et al. (2010) and previously found in the lake (gkelis and zaoutsos, 2014), however the neostx/ stx frequency (100% of the samples analysed by lc-ms/ms) is one of the highest ever reported. atx-a occurrence was reported for the first time in a greek freshwater in a previous study by dimitrakopoulos et al. (2010) although it did not specifically report which freshwaters were sampled. the atx-a concentrations we found here are similar to those recently reported by toporowska et al. (2016). mc concentrations during the monitoring period were in the range reported previously (papadimitriou et al., 2013), much lower than the ones found at off-shore stations (gkelis and zaoutsos, 2014). in our study cyn and stx was detected in blooms consisting of mixed populations of a. elenkinii, sph. aphanizomenoides, and c. raciborskii, in line with our previous findings in lake karla (gkelis and zaoutsos, 2014). recently it was shown that aphanizomenon seems to be the main cyanobacterial genus responsible for the production of cyn in polish lakes (mankiewicz-boczek et al., 2012). evidences point to aph. gracile as the stx producer in europe (ballot et al., 2010; cirés et al., 2014). gkelis and zaoutsos (2014) suggested that c. raciborskii and aphanizomenon (aph. cf. flos-aquae) could be the possible stx producers in greece. atx-a has been found in dolichospermum (anabaena) and aphanizomenon populations worldwide (testai et al., 2016). the suspected cyn or atx-a production of aph. flos-aquae or mc production of sph. aphanizomenoides is still uncertain (cirés and ballot, 2016). further studies involving strains and/or cyanotoxins gene sequences are needed to identify the cyanotoxins in lake karla and kalamaki reservoir, as in all cases mixed cyanobacteria populations co-occurred with a mixture of cyanotoxins. it has been well documented that environmental factors such as temperature, ph, dissolved oxygen, and nutrient availability play an important role in regulating the structure and distribution of phytoplankton communities in lakes (dolman et al., 2012; paerl and paul, 2012), whereas, their influence on cyanotoxin production is more complex (boopathi and ki, 2014). our results indicated that the intracellular mc and cyn concentration was linked to water temperature and srp concentrations. the positive correlation of mcs to temperature, and srp, found also recently in lake pamvotis (gkelis et al., 2014) can be explained by the fact that in high concentrations of phosphorus (o’ neil et al., 2012), or nitrogen or joint nitrogen-phosphorus (dolman et al., 2012) hepatotoxic strains produce more mcs. however, given the very high level of srp throughout the study we cannot conclude that srp has been a determining factor. to date, very few studies have investigated the relationship between cyn concentration and environmental factors potentially affecting its production. experimental studies were focused, e.g., on effects of temperature (saker and griffiths, 2000), nitrogen (saker and neilan, 2001), or light (dyble et al., 2006; bormans et al., 2014). our work showed a positive correlation of cyn to temperature in line with kokociński et al. (2013); earlier experimental studies by saker and griffiths (2000) and preußel et al. (2009) found a reduction of cyn production at temperatures over 20°c and 25°c, respectively. these somewhat contrasting results could be due to the different cyn producers examined and different regulatory mechanisms of cyn-producing strains involved as wiedner et al. (2008) point out. in the present study, the phytoplankton biomass (dominated by cyanobacteria) and the mcs concentrations were well above the who guidance level 2 (who, 2003) for recreational waters, posing a moderate health risk throughout the year. in some cases (warm period in lake karla and february 2013 in kalamaki reservoir), biomass values exceeded also the guidance level 3, without, however, the mcs concentration being equally high. nevertheless, given the co-occurrence of multiple cyanotoxins in lake karla, documented in our study, and the fish mortalities (oikonomou et al., 2012) and mcs accumulation in fish (papadimitriou et al., 2013) it should be considered that guidance level 3 health risks (who, 2003) are possible. therefore, the recommended actions (who, 2003), including possible prohibition of watercontact activities, public health follow-up investigation, and informing of the relevant authorities, should be taken. it should be noted that the frequency of recurrent waterbloom phenomena, especially during the warm period, could be much higher than the sampling of this monitoring program, thus a quick and efficient water-bloom early warning system should be implemented. an overall assessment of the monitoring in lake karla and kalamaki reservoir indicates that, as in other greek lakes (latinopoulos et al., 2016), multiple stressors act synergistically in degrading their ecological status. while the diffuse pollution from agriculture is the single most important source of pollution in most european lakes nowadays, greek lakes still face point pollution, particularly by nutrients (latinopoulos et al., 2016). in lake karla, nutrient load combining both diffuse and point pollution is 8200 kg day–1 for n and 870 kg day–1 for p, res. gkelis et al.48 spectively (egy, 2013). thus, a drastic reduction of nutrient load should be achieved ‘at source level’ and before entering the riparian zone. controlling nutrients remains the basis for managing blooms, no matter which phytoplankton functional type dominates (mantzouki et al., 2016), therefore the decrease of fertilizer level and the control of the point pollution sources are considered as the only feasible measures. furthermore, in lake karla the continuous water level decline from the suggested ‘lower ecological water level’ (i.e., 46.4 m asl) causes accumulation of nutrients which is strengthening the eutrophication. we have not evidence yet about the extent of the nutrient’s re-suspension effect through the sediment but it is likely to happen (scheffer, 2004; christophoridis et. al, 2006) since the bottom experienced a long fertilization period during the dryness time. according to the present nutrient and phytoplankton biomass levels, it becomes clear that lake karla is a eutrophic system, with apparent signals of hypertrophication during the warm period (chamoglou et al., 2014; theologou et al., 2016). the absence of any outlet, and thus of any flushing process, leads to a high-water residence time, strengthening the eutrophic conditions. in shallow mediterranean lakes, nutrient inputs from the catchment occur mainly in winter-spring due to high precipitation, whereas when there is no outflow, they act as ‘nutrient sinks’ (chamoglou et al., 2014). the reduction of the residence time by regulating the annual timing of the inflows and outflows could certainly aim at the improvement of water quality in lake karla. this is considered as an emergent issue concerning the future management process. conclusions the ecological potential of lake karla and kalamaki reservoir was less than good based on the biological qualitative elements (benthic macroinvertebrates and mainly phytoplankton) and the co-occurrence of multiple cyanotoxins, implying an intense deterioration of the lake water quality. our first observations for this deterioration, subject to further investigation, point to: i) the quality of water inflowing from the ditches to both systems, connected to pinios river; ii) the quantitative management of the lake water (i.e., the irrigation/drainage system of karla’s basin via complex networks and interaction with fields under intensive agriculture); and iii) perhaps the absence of management of the reed bed in the littoral zone of the lake. the new lake is in the first stage of water filling and its wetland systems are not yet fully functional in order to act as an artificial outflow and reduce incoming organic load by abducting part of the lake’s water through the irrigation network. a permanent quality and quantity monitoring system to ensure immediate intervention when incoming pollution loads are large will help habitat protection in lake karla and kalamaki reservoir. according to the national management plan (egy, 2013), the completion of works rendering the hydraulic balance of karla and the rationalisation in the consumption of irrigatory water would help the improvement of the water quality and further the ecological status. acknowledgments this study was partially funded by the research program ‘monitoring of water quality parameters’ funded by the management body of ecodevelopment area of karla, mavrovouni, kefalovriso, velestino, stefanovikio (aristotle university of thessaloniki research committee contract no 89623) and managed by omikron ltd. the authors gratefully acknowledge cyanocost-cost es 1105 for sharing of knowledge and networking. disclaimer (t. kaloudis): the views expressed in this manuscript do not necessarily reflect the views of eydap sa. author contribution: sg, ml, ik, conceived and designed the study; sg, ic, mp, collected the samples; ic, performed the nutrients’ analyses; sg, mp, performed the phytoplankton, molecular, and elisa analyses; cn, performed the benthic invertebrates’ analysis; s-kz, cc, km, tmt, tk, ah, performed the lc-ms/ms analyses; sg, mp, visualized the data; sg manuscript drafting. all authors contributed in editing and/or reviewing the manuscript. references ananiadis ci, 1956. limnological study of lake karla. b. inst. ocean. 1083:1-19. armitage pd, hogger j, 1994. invertebrates ecology and methods of survey, p. 85-97, 151-159. in: n. holmes, d. ward and p. jose, the new rivers and wildlife handbook. rspb sandy bedfordshire uk: 426 pp. atashpaz s, khani s, barzegari a, barar j, vahed sz, azarbaijani r, omidi y, 2010. a robust universal method for extraction of genomic dna from bacterial species. microbiology 79:538-542. ballot a, fastner j, wiedner c, 2010. paralytic shellfish poisoning toxin-producing cyanobacterium aphanizomenon gracile in northeast germany. appl. environ. microbiol. 76:1173-1180. bartram j, burch m, falconer ir, jones g, kuiper-goodman t, 1999. situation assessment, planning and management, pp. 179-209. in: i. chorus and j. bartram (eds.), toxic cyanobacteria in water. 1st ed. world health organization, e. & f.n. spon. beklioglu m, romo s, kagalou i, quintana x, becares e, 2007. state of the art in the functioning of shallow mediterranean lakes: workshop conclusions. hydrobiologia 584:317-26. water quality and cyanotoxins in the reconstructed lake karla 49 berry jp, lind o, 2010. first evidence of “paralytic shellfish toxins” and cylindrospermopsin in a mexican freshwater system, lago catemaco, and apparent bioaccumulation of the toxins in “tegogolo” snails (pomacea patula catemacensis). toxicon 55:930-938. boopathi t, ki j-s, 2014. impact of environmental factors on the regulation of cyanotoxin production. toxins 6:1951-1978. bormans m, lengronne m, brient l, duval c, 2014. cylindrospermopsin accumulation and release by the benthic cyanobacterium oscillatoria sp. pcc 6506 under different light conditions and growth phases. bull. environ. contam. toxicol. 92:243-247. brient l, lengronne m, bormans m, fastner j, 2009. first occurrence of cylindrospermopsin in freshwater in france. environ. toxicol. 24:415-420. chamoglou m, papadimitriou t, kagalou i, 2014. keys-descriptors for the functioning of a mediterranean reservoir: the case of a new lake karla-greece. environ. process. 1:127-135. christophoridis c, fytianos k 2006. conditions affecting the release of phosphorus from surface lake sediments. j. environ. qual. 35:1181-1192. cirés s, ballot a, 2016. a review of the phylogeny, ecology and toxin production of bloom-forming aphanizomenon spp. and related species within the nostocales (cyanobacteria). harmful algae 54:21-43cirés s, wörmer l, ballot a, agha r, wiedner c, velázquez d, casero mc, quesada a, 2014. phylogeography of cylindrospermopsin and paralytic shellfish toxin-producing nostocales cyanobacteria from mediterranean europe (spain). appl. environ. microbiol. 80:1359-1370. codd ga, lindsay j, young fm, morrison lf, metcalf js, 2005. harmful cyanobacteria: from mass mortalities to management measures, p. 1-23. in: j. huisman, h.c.p. matthijs, p.m. visser (eds.), harmful cyanobacteria. springer, dordrecht. cook cm, vardaka e, lanaras t, 2004. toxic cyanobacteria in greek freshwaters, 1987–2000: occurrence, toxicity, and impacts in the mediterranean region. acta hydroch. hydrob. 32:107-124. cooke gd, welch eg, peterson sa, newroth pr, 1986. limnology, lake diagnosis and selection of restoration methods, p. 9-46. in: g.d. cooke, e.g. welch, s.a. peterson and p.r. newroth (eds.), lake and reservoir restoration. butterworth publ., stoneham. costa ias, azevedo smfo, senna pac, bernardo rr, costa sm, chellappa nt, 2006. occurrence of toxin-producing cyanobacteria blooms in a brazilian semiarid reservoir. braz. j. biol. 66:211–219. dimitrakopoulos i, kaloudis t, hiskia a, thomaidis ns, koupparis ma, 2010. development of a fast and selective method for the sensitive determination of anatoxin-a in lake waters using liquid chromatography–tandem mass spectrometry and phenylalanine-d 5 as internal standard. anal. bioanal. chem. 397:2245-2252. dokulil mt, teubner k, 2000. cyanobacterial dominance in lakes. hydrobiologia 438:1-12. dokulil mt, 2016. vegetative survival of cylindrospermopsis raciborskii (cyanobacteria) at low temperature and low light. hydrobiologia 764:241-247. dolman am, rücker j, pick fr, fastner j, rohrlack t, mischke u, wiedner c, 2012. cyanobacteria and cyanotoxins: the influence of nitrogen and phosphorus. plos one 7:e38757. dyble j, tester pa, litaker rw, 2006. effects of light on cylindrospermopsin production in the cyanobacterial hab species cylindrospermopsis raciborskii. afr. j. mar. sci. 28:309-312. egy, 2013. [river basin management plan of thessaly].[report in greek]. accessed on: 01/05/2017. available from: http:// wfd.ypeka.gr/pdf/gr08_sxedio_diaxeirisis%20neron_ ypogegrammeno.pdf en 27828-1994. water quality; methods of biological sampling; guidance on handnet sampling of aquatic benthic macro-invertebrates (iso 7828, 1985). fergusson km, saint cp, 2003. multiplex pcr assay for cylindrospermopsis raciborskii and cylindrospermopsin-producing cyanobacteria. environ. toxicol. 18:120-125. gkelis s, harjunpää v, lanaras t, sivonen k, 2005. diversity of hepatotoxic microcystins and bioactive anabaenopeptins in cyanobacterial blooms from greek freshwaters. environ. toxicol. 20:249-256. gkelis s, lanaras t, sivonen k, 2015. cyanobacterial toxic and bioactive peptides in freshwater bodies of greece: concentrations, occurrence patterns, and implications for human health. mar. drugs 13:6319-6335. gkelis s, papadimitriou t, zaoutsos n, leonardos i, 2014. anthropogenic and climate-induced change favors toxic cyanobacteria blooms: evidence from monitoring a highly eutrophic, urban mediterranean lake. harmful algae 39:322-333. gkelis s, zaoutsos n, 2014. cyanotoxin occurrence and potentially toxin producing cyanobacteria in freshwaters of greece: a multi-disciplinary approach. toxicon 78:1-9. graham j, loftin k, meyer m, ziegler a, 2010. cyanotoxin mixtures and taste-and-odor-compounds in cyanobacterial blooms from the midwestern united states. environ. scien. technol. 44:7361-7368. hisbergues m, christiansen g, rouhiainen l, sivonen k, borner t, 2003. pcr-based identification of microcystinproducing genotypes of different cyanobacterial genera. arch. microbiol. 180:402-410. iso 7828, 1985. water quality methods of biological sampling – guidance on handnet sampling of aquatic benthic macroinvertebrates. international organization for standardization. jeppesen e, søndergaard m, meerhoff m, lauridsen tl, jensen jp, 2007. shallow lake restoration by nutrient loading reduction some recent findings and challenges ahead. hydrobiologia 584:239-252. jrc european commission, 2009. water framework directive intercalibration technical report. part 2: lakes. joint research centre, european commission. jungblut ad, neilan ba, 2006. molecular identification and evolution of the cyclic peptide hepatotoxins, microcystin and nodularin, synthetase genes in three orders of cyanobacteria. arch. microb. 185:107-114. kagalou i, 2010. classification and management issues of greek lakes under the european water framework directive: a dpsir approach. j. environ. monitor. 12:2207-2215. kagalou i, papastergiadou e, leonardos i, 2008. long term changes in the eutrophication process in a shallow mediterranean lake ecosystem of w. greece. response after the reduction of external load. j. environ. manag. 87:497-506. kellmann r, mihali tk, jeon yj, pickford r, pomati f, neilan s. gkelis et al.50 ba, 2008. biosynthetic intermediate analysis and functional homology reveal a saxitoxin gene cluster in cyanobacteria. appl. environ. microb. 74:4044-4053. kokociński m, dziga d, spoof l, stefaniak k, jurczak t, mankiewiczboczek j, meriluoto j, 2009. first report of the cyanobacterial toxin cylindrospermopsin in the shallow, eutrophic lakes of western poland. chemosphere 74:669-675. kokociński m, mankiewiczboczek j, jurczak t, spoof l, meriluoto j, rejmonczyk e, hautala h, vehniäinen m, pawełczyk j, soininen j, 2013. aphanizomenon gracile (nostocales), a cylindrospermopsin producing cyanobacterium in polish lakes. environ. sci. pollut. res. 20:52435264. kolzau s, wiedner c, rücker j, köhler j, köhler a, dolman am, 2014. seasonal patterns of nitrogen and phosphorus limitation in four german lakes and the predictability of limitation status from ambient nutrient concentrations. plos one 9:e96065. komárek j, 2013. [cyanoprokaryota 3. teil: heterocytous genera]. in: b. büdel, g. gärtner, l. krienitz and m. schagerl (eds.), [süswasserflora von mitteleuropa, freshwater flora of central europe].[book in german]. springer spektrum berlin: 1130 pp. komárek j, anagnostidis k, 1999.[ cyanoprokaryota 1. teil: chroococcales]. in: h. ettl, g. gärtner, h. heynig, and d. mollenhauer (eds.), [süsswasserflora von mitteleuropa 19/1].[book in german]. gustav fischer verlag, jena: 548 pp komárek j, anagnostidis k, 2005. [cyanoprokaryota 2. teil: oscillatoriales]. in: h. ettl, g. gärtner, h. heynig, and d. mollenhauer (eds.), [süsswasserflora von mitteleuropa 19/1].[book in german]. gustav fischer verlag, jena: 759 pp. lakes network, 2016. lake paralimni. accessed on: 02/24/2017. available from: http://lakesnetwork.eu/english/lakes/lake-paralimni-2/ latinopoulos d, ntislidou c, kagalou i, 2016. multipurpose plans for the sustainability of the greek lakes: emphasis on multiple stressors. environ. process. 3:589. loukas a, mylopoulos n, vasiliades l, 2007. a modelling system for the evaluation of water resources management scenarios in thessaly, greece. water resour. manage. 21:1673-702. mankiewicz-boczek j, kokocinski m, gagala i, pawelczyk j, jurczak t, dziadek j, 2012. preliminary molecular identification of cylindrospermopsin-producing cyanobacteria in two polish lakes (central europe). microbiol. lett. 326:173-179. mantzouki e, visser pm, bormans m, ibelings bw. 2016. understanding the key ecological traits of cyanobacteria as a basis for their management and control in changing lakes. aquat. ecol. 50:333-350. mischke u, nixdorf b, 2003. equilibrium phase conditions in shallow german lakes: how cyanoprokaryota species establish a steady state phase in late summer. hydrobiologia 502:123-132. mischke u, 2003. cyanobacteria associations in shallow polytrophic lakes: influence of environmental factors. acta oecologica 24:s11-s23. moustakagouni m, michaloudi e, sommer u, 2014. modifying the peg model for mediterranean lakes – no biological winter and strong fish predation. freshwater biol. 59:11361144. moustaka-gouni m, hiskia a, genitsaris s, katsiapi m, manolidi k, zervou s-k, christophoridis c, triantis tm, kaloudis t, orfanidis s, 2016. first report of aphanizomenon favaloroi occurrence in europe associated with saxitoxins and a massive fish kill in lake vistonis, greece. mar. freshwater res. doi: http://dx.doi.org/10.1071/mf16029 nonneman d, zimba p, 2002. a pcr-based test to assess the potential for microcystin occurrence in channel catfish production ponds. j. phycol. 38:230-234. o’neil jm, davis tw, burford ma, gobler cj, 2012. the rise of harmful cyanobacteria blooms: the potential roles of eutrophication and climate change. harmful algae 14:313-334. oikonomou a, katsiapi m, karayanni h, moustaka-gouni m, kormas, k, 2012. plankton microorganisms coinciding with two consecutive mass fish kills in a newly reconstructed lake. sci. world j. 2012:1-14. ouahid y, pérez-silva g, campo ffd, 2005. identification of potentially toxic environmental microcystis by individual and multiple pcr amplification of specific microcystin synthetase gene regions. environ. toxicol. 20:235-242. padisák j, borics g, grigorszky i, soróczki-pintér é, 2006. use of phytoplankton assemblages for monitoring ecological status of lakes within the water framework directive: the assemblage index. hydrobiologia, 553:1-14. paerl hw, 2014. mitigating harmful cyanobacterial blooms in a humanand climatically-impacted world. life. 4:988-1012. paerl hw, hall ns, calandrino es, 2011. controlling harmful cyanobacterial blooms in a world experiencing anthropogenic and climatic-induced change. sci. total environ. 409:1739-1745. paerl hw, otten gt, 2013. harmful cyanobacterial blooms, causes, consequences, and controls. microb. ecol. 65:995– 1010. paerl hw, paul v, 2012. climate change: links to global expansion of harmful cyanobacteria. water res. 46:1349-1363. papadimitriou t, katsiapi m, kormas ka, moustaka-gouni m, kagalou i, 2013. artificially-born “killer” lake: phytoplankton based water quality and microcystin affected fish in a reconstructed lake. sci. total environ. 452-453:116-124. preußel k, wessel g, fastner j, chorus i, 2009. response of cylindrospermopsin production and release in aphanizomenon flos-aquae (cyanobacteria) to varying light and temperature conditions. harmful algae 8:645-650. reynolds cs, huszar v, kruk c, naselli-flores l, melo s, 2002. towards a functional classification of the freshwater phytoplankton. j. plankton res. 24:417-428. romo s, miracle mr, villena mj, rueda m, ferriol c, vicente e, 2004. mesocosm experiments on nutrient and fish effects on shallow lake food webs in a mediterranean climate. freshwater biol. 49:1593-607. rücker j, stüken a, nixdorf b, fastner j, chorus i, wiedner c, 2007. concentrations of particulate and dissolved cylindrospermopsin in 21 aphanizomenon-dominated temperate lakes. toxicon 50:800-809. rzymski p, poniedziałek b, 2014. in search of environmental role of cylindrospermopsin: a review on global distribution and ecology of its producers. water research 66:320-337. saker ml, griffiths dj, 2000. effects of temperature on growth and cylindrospermopsin content of seven isolates of cylindrospermopsis raciborskii (nostocales, cyanophyceae) from water bodies in northern australia. phycologia 39:349-354. water quality and cyanotoxins in the reconstructed lake karla 51 saker ml, neilan ba, 2001. varied diazotrophies, morphologies and toxicities of genetically similar isolates of cylindrospermopsis raciborskii (nostocales, cyanophyceae) from water bodies in northern australia. appl. environ. microbiol. 99:749-757. scheffer m, 2004. ecology of shallow lakes. springer science+business media, dordrecht: 378 pp. schembri ma, neilan ba, saint cp, 2001. identification of genes implicated in toxin production in the cyanobacterium cylindrospermopsis raciborskii. environ. toxicol.16:413-421. šejnohová l, maršálek b, 2012.microcystis, p. 195-221. in: b.a. whitton (ed.), ecology of cyanobacteria. ii. their diversity in space and time. springer, dordrecht. sidiropoulos p, papadimitriou t, stabouli z, loukas a, mylopoulos n, kagalou i, 2012. past, present and future concepts for conservation of the re-constructed lake karla (thessaly-greece). fresenius environ. bull. 21:3027-34. testai e, scardala s, vichi s, buratti fm, funari e, 2016. risk to human health associated with the environmental occurrence of cyanobacterial neurotoxic alkaloids anatoxins and saxitoxins. crit. rev. toxicol. 46:385-419. theologou i, kagalou i, papadopoulou mp, karantzalos k, 2016. multitemporal mapping of chlorophyll-α in lake karla from high resolution multispectral satellite data. environ. process. 3:681. timm t, 2009. a guide to the freshwater oligochaeta and polychaeta of northern and central europe. lauterbornia 66: 1-235. toporowska m, pawlik-skowrońska b, kalinowska r, 2016. mass development of diazotrophic cyanobacteria (nostocales) and production of neurotoxic anatoxin-a in a planktothrix (oscillatoriales) dominated temperate lake. water air soil pollut. 227:321. utermöhl h, 1958. [zur vervollkommung der quantitativinen phytoplankton-methodik].[article in german]. int. ver. theor. angew. limnol. 9:1-38. vasconcelos v, martins a, vale m, antunes a, azevedo j, welker m, lopez o, montejano g, 2010. first report on the occurrence of microcystins in planktonic cyanobacteria from central mexico. toxicon 56:425-431. villena mj, romo s, 2003. phytoplankton changes in a shallow mediterranean lake (albufera of valencia, spain) after sewage diversion. hydrobiologia. 506:281-287. visser pm, ibelings bw, bormans m, huisman j. 2016. artificial mixing to control cyanobacterial blooms: a review. aquat. ecol. 50:423-441. who (world health organization), 2003. guidelines for safe recreational water environments. 1. coastal and fresh waters. who, geneva. wiederholm t, 1983. chironomidae of the holarctic region. keys and diagnoses. part i. larvae: entomol. scand. supplement, 1-457. wiederholm t, 1986. chironomidae of the holarctic region. keys and diagnoses. part ii. pupae: entomol. scand. supplement, 1-482. wiedner c, rücker j, fastner j, chorus i, nixdorf b, 2008. seasonal dynamics of cylindrospermopsin and cyanobacteria in two german lakes. toxicon 52:677–786. zervou s-k, christophoridis c, kaloudis t, triantis tm, hiskia a, 2017. new spe-lc-ms/ms method for simultaneous determination of multi-class cyanobacterial and algal toxins. j. hazard. mater. 323:56-66. layout 1 advances in oceanography and limnology, 2016; 7(1): 36-50 article doi: 10.4081/aiol.2016.5791 introduction water temperature in lakes is governed by a complex heat budget resulting from the combination of different heat flux components that are mainly exchanged between the lake and the atmosphere. water temperature is the primary driver of vertical stratification in lakes, thus it significantly affects transport of mass (including nutrients and dissolved oxygen), energy, and momentum within the water column. it crucially controls several physical (e.g., thermal stratification, mixing processes), geochemical (e.g., chemical reaction rates, oxygen solubility), and ecological (e.g., metabolism, growth, and reproduction of organisms) processes, with considerable influences on the overall lake water quality, ecosystem functioning, and community composition (wetzel, 2001; gallina et al., 2013). it is therefore evident that any significant changes in water temperature may lead to alterations in the thermal regime of the lake and in the community structure of many freshwater habitats (winder and sommer, 2012; de senerpont domis et al., 2013; schabhüttl et al., 2013; butcher et al., 2015), with possible modifications of the biochemical compositions of some algae species (flaim et al., 2014). this is particularly relevant considering that lakes have been demonstrated to be highly sensitive to changes in environmental conditions (adrian et al., 2009; o’reilly et al., 2015). in the light of the above considerations, large efforts have been directed towards the development of models able to predict water temperature, with a particular attention to lake surface temperature (lst). several models of different type and complexity have been proposed to simulate water temperature, ranging from simple regression models (mccombie, 1959; webb, 1974; livingstone and lotter 1998; kettle et al., 2004; sharma et al., 2008) to more complex process-based numerical models (perroud et al., 2009; martynov et al., 2010; thiery et al., 2014). regressions models are attractive because they require little information, usually only air temperature, but generally they are not able to address some fundamental processes (e.g., the prediction of lake surface temperature using the air2water model: guidelines, challenges, and future perspectives sebastiano piccolroaz department of civil, environmental and mechanical engineering, university of trento, via mesiano 77, i-38123,trento, italy corresponding author: s.piccolroaz@unitn.it abstract water temperature plays a primary role in controlling a wide range of physical, geochemical and ecological processes in lakes, with considerable influences on lake water quality and ecosystem functioning. being able to reliably predict water temperature is therefore a desired goal, which stimulated the development of models of different type and complexity, ranging from simple regression-based models to more sophisticated process-based numerical models. however, both types of models suffer of some limitations: the first are not able to address some fundamental physical processes as e.g., thermal stratification, while the latter generally require a large amount of data in input, which are not always available. in this work, lake surface temperature is simulated by means of air2water, a hybrid physically-based/statistical model, which is able to provide a robust, predictive understanding of lst dynamics knowing air temperature only. this model showed performances that are comparable with those obtained by using process based models (a root mean square error on the order of 1°c, at daily scale), while retaining the simplicity and parsimony of regression-based models, thus making it a good candidate for long-term applications. the aim of the present work is to provide the reader with useful and practical guidelines for proper use of the air2water model and for critical analysis of results. two case studies have been selected for the analysis: lake superior and lake erie (usa). these are clear and emblematic examples of a deep and a shallow temperate lake characterized by markedly different thermal responses to external forcing, thus are ideal for making the results of the analysis the most general and comprehensive. particular attention is paid to assessing the influence of missing data on model performance, and to evaluating when an observed time series is sufficiently informative for proper model calibration or, conversely, data are too scarce thus leading to the risk of overfitting. the final aim of the work is to facilitate the use of the model also by scientists that do not necessarily have a solid background on modeling or physics. this work is also an attempt to foster the communication and interaction among colleagues of a branch of science, limnology, which suffer of significant fragmentation. this is summarized in the future perspectives and challenges concerning potential improvements of the air2water, with a particular emphasis on possible cross-sectoral applications key words: lake surface temperature; air2water; air temperature, thermal response; temperature modeling. received: february 2016. accepted: march 2016. no nco mm er cia l u se on ly 37s. piccolroaz effect of thermal stratification) and their use may be questionable especially when it is necessary to extrapolate temperature values beyond the limits of the measured time series, as is typically the case in climate change studies. on the other hand, deterministic models are designed to provide an exhaustive description of the thermal behaviour of the lake, but they require detailed time series of meteorological variables, which are not always available for long periods and with a sufficient accuracy. in order to overcome the limitations of traditional approaches, piccolroaz et al. (2013) recently developed air2water, a hybrid physically-based/statistical model, which is able to provide a robust, predictive understanding of lst dynamics knowing air temperature only. the hybrid formulation of the air2water model combines a physically based derivation of the governing equation with a stochastic calibration of the parameters. in this way, the information contained in the data is transferred directly to model parameters, whose calibrated values can provide significant information as to how the real system behaves (thanks to the physical-based structure of the model). the underlying rationale behind the development of this model is to take advantage of the fact that the governing laws of physics are generally well understood, to introduce opportune simplifications while retaining all the fundamental processes (and their physical meaning) involved. the purpose is to minimize data requirements and computational effort, which still represent the most common limitations, and to develop a as simple as possible but not simpler (citing a famous quote by albert einstein) mathematical tool able to provide a reliable description of a natural phenomenon on the basis of the data that are available. the model has been successfully tested considering lakes characterized by different morphometric characteristics and using different sources of data (see e.g., toffolon et al., 2014a, who applied the air2water model to 14 different lakes in the temperate region: 7 located in north america, 6 in europe, and 1 in asia). in all cases, air2water performed similarly to more complex process-based models (i.e., rmse on the order of 1°c for daily temperatures), even though these latter models generally require a much larger amount of information. the model has been shown to satisfactorily capture seasonal variations and inter-annual dynamics of lst, and to provide key information to investigate the role of stratification in controlling the thermal response of lakes (piccolroaz et al., 2015a). this work provides the reader with practical guidelines for proper use of the air2water model and for critical analysis of results, with the final goal of facilitating the use of the model by scientists that do not necessarily have a solid background on modelling or physics. however, this work should not be considered simply as a collection of best practices, but also as an attempt to foster communication among colleagues from different disciplines with a common interest in aquatic science. the reader will find answers to questions like: what is the meaning of model parameters, how are they derived, and how should we select their a priori range of variation?; what is the maximum allowable percentage of missing data to obtain reliable results?; how long should the calibration period be?; what version of the model should be used?; does lake depth affect model performance?. particular attention is given to analysing the effects of data scarcity on model performance in modelling lst. finally, future directions and perspectives concerning possible improvements of the air2water model are discussed, with a particular emphasis on cross-sectoral applications. methods study sites and available data the air2water model is applied to two lakes characterized by significantly different morphological and thermal characteristics: lake superior and lake erie (usa) (fig. 1). lake superior is the largest, deepest, and most northern of the great lakes, while lake erie is the smallest, shallowest, and most southern of the two lakes in this study (tab. 1). long-term data of air and surface water temperature are available from different sources. in this work the following sources of data are used: glsea daily lst retrieved from satellite imagery (i.e., skin temperature) provided by national oceanic and atmospheric administration (noaa) great lakes environmental research laboratory (glerl, webpage: http://www. glerl.noaa.gov/glsea/asc_ 1024/) and derived from noaa polar-orbiting satellites equipped with avhrr sensors, and daily air temperature at 2 meters above ground from era-interim reanalysis (provided by the european centre for medium-range weather forecasts, ecmwf and downloaded from http://apps.ecmwf.int/datasets/data/interim-full-daily/levtype=sfc/). both datasets cover the 20year period 1995-2014 and contain spatially distributed data (with resolution equal to about 1.3 km and 80 km in the two cases, respectively). the data have been postprocessed in order to evaluate lake-average values, i.e., temperature values have been aggregated at the lake scale. moreover, in order to allow the analyses, as presented in the results section, the missing data in the lst series have been replaced by interpolation with a moving average filter of 10 days. fig. 1 shows the typical annual cycles of air and water temperature for the two lakes, and suggests the existence of markedly different thermal behaviours: i) due to the higher latitude, air temperature over lake superior is generally colder than for lake erie (annual mean, minimum, and maximum equal to about 3.9°c vs 9.6°c, -13.9°c vs no nco mm er cia l u se on ly 38 the air2water model: guidelines, challenges, and perspectives -6.4°c, and 17.8°c vs 23.5°c, for the two lakes respectively) and the maximum air temperature occurs later (beginning of august vs middle of july); ii) the amplitude of the phase lag (hysteresis) between air and water temperature is more evident for lake superior than for lake erie indicating a larger thermal inertia due to the larger water volume; iii) consequently, the onset of direct thermal stratification (i.e., when tw≥4°c)) in lake superior occurs later in the year (end of may vs middle of april) as well as the period of maximum stratification (i.e., when tw is maximum; end vs beginning of august); iv) the shape of lst annual cycle deviates from the nearly sinusoidal pattern of air temperature in lake superior, contrary to what happens in the case of lake erie; and v) lake superior is generally colder than lake erie (mean annual lst equal to 6.5°c and 11.3°c in the two lakes, respectively). the choice of these two case studies is not only motivated by the large amount of high quality and freely available data, but also, and more importantly, by the fact that they are good examples of deep and shallow temperate lakes characterized by markedly different thermal responses to external forcing. this requisite is certainly of major importance in order to write as much as possible exhaustive and generally valid guidelines for best practices around the use of the air2water model. the air2water model the air2water model is based on a lumped heat budget of the surface volume of the lake at daily time scale, and is derived from the following volume-integrated heat equation: (eq. 1) from which the variation of water temperature (tw) in time (t: hereafter expressed in days) is directly dependent on the product between the heat flux into the upper water volume (hnet) and the surface area of the lake (a), and inversely dependent on the surface volume of water involved in the heat exchange with the atmosphere (vs: hereafter also referred to as the reactive volume), density (ρ) (1000 kg m–3), and specific heat capacity at constant pressure (cp) (4186 j kg–1 °c–1). hnet can be expressed as the combination of several contributions entering and exiting the upper water volume (vs) (see fig. 2 for a schematic, and supplementary material a for details), which are primarily controlled by: the net shortwave (hs) and longwave (ha) radiation actually absorbed by the surface volume (i.e., accounting for water reflectivity), the longwave radiation emitted from the lake (hw), the latent heat flux due to evaporation and condensation (hl), and the sensible heat flux due to convection (hc). heat flux due to precipitation, the heat exchanged with inlets/outlets, and the heat exchanged between surface volume and deep water or sediments can be considered as insignificant fig. 1. geographical location of lake superior and lake erie in the great lakes region and in north america. typical annual cycles (averaged over the period 1995 to 2014) of air and water temperature for the two lakes. tab. 1. main morphological characteristics of the investigated lakes. volume (km3) surface area (km3) maximum depth (m) average depth (m) geographic coordinates lake superior 12,000 82,100 406 147 47.7°n 87.5°w lake erie 480 25,667 64 19 42.2°n 81.2°w no nco mm er cia l u se on ly 39s. piccolroaz factors, and are not explicitly included in the formulation of air2water. however, their contribution is indirectly accounted for in the calibration of parameters. following livingstone and padisák (2007), air temperature can be considered as a proxy for the integrated effect of the external forcing, and it can be assumed, together with lst, as the key factor controlling the heat balance of the surface layer of the lake. this is the central concept of the air2water model. in particular, hnet is included in a linear form obtained by taylor expansion in terms of both air (ta) and water (tw) temperatures, as follows: (eq. 2) whereand t̄w are reference values (e.g., long term averages of ta and tw, respectively), and hnet,0=hnet | t̄a , t̄w is the part of hnet that is independent on air and water temperatures. in general, however, hnet,0 can vary in time. as a first approximation, this is accounted for by defining hnet,0 as the sum of a constant value and a sinusoidal function of time with a period of 1 year, the latter term summarizing, albeit in a simplified form, the combined effect due to the variability of all meteorological variables other than air temperature (e.g., solar radiation, wind speed, air humidity, cloudiness) at annual time scale. equation (1) can be therefore rewritten as follows: (eq. 3) where the definition of parameters âi,i=1, 2, 3, 5, 6 can be derived from equation (2) once the single heat flux terms are evaluated through suitable empirical relationships (martin and mccutcheon, 1998). refer to supplementary material a for details about the linearization hnet of , and the definition of parameters âi. by introducing the dimensionless ratio δ=vs /vr (which can be also interpreted as the ratio between the average depth of the surface layer ds=vs /a and that of the reference layer dr=vr /a.), eq. (3) can be rewritten as the following ordinary differential equation, representing the full version of the air2water model: (eq. 4) where parameters ai,i=1, 2, 3, 5 are defined as ai=âia/(vr ρcp)=âi/(dr ρcp). in this form, the geometrical characteristics of the lake (surface area, volume, and depth) are not required to be explicitly specified, since are implicitly accounted for in the model parameters ai, which require calibration. in order to ensure proper model calibration excluding unrealistic solutions, the model parameters are allowed to vary within a physically plausible range, which can be easily estimated knowing (even approximately) the mean depth of the lake, as will be thoroughly discussed in the results section. equation (4) is numerically integrated with a daily time step (i.e., dt=1 day; see also the methods section for further details). finally, in order to account for the significant seasonal variability of the reactive volume as a consequence of thermal stratification, piccolroaz et al. (2013) assumed that the dimensionless ratio (δ) is a function of the difference between lst and a reference value of the deep water temperature (th), through the following empirical relationship: (eq. 5) where th can be assumed to be 4°c for dimictic lakes, and the minimum or maximum water temperature for warm and cold monomictic lakes, respectively, and a4, a7, and a8 are model parameters. from the first formula in equation (5) it is easy to see that the dimensionless ratio δ is theoretically defined in a range from 0 to 1, with δ decreasing for increasing thermal stratification (here represented by the difference tw–th), thus mimicking the fact that the surface water volume affected by the surface heat budget gets progressively thinner. conversely, δ=1 when the lake is isothermal (i.e., tw–th), suggesting that the reference volume can be interpreted as the maximum water volume involved in the heat exchange with the atmosphere during the year. the same considerations apply to the second formula in equation (5), which is valid when the lake is inversely stratified (i.e., when ). in this case, however, the possible effect of heat flux reduction due to ice cover is also included by a fictitious increase of the effective volume (see the second term on the right-hand side). in order to simulate ice formation at the surface, a lower bound is imposed on by introducing a threshold value. this threshold is generally 0°c when the water temperature is measured close to the surface, but it can be higher when temperature is measured at deeper depths. despite being simple, the parameterization of δ presented in equation (5) is suitable to reproduce fig. 2. main heat fluxes involved in the heat budget of the surface layer. see supplementary material a for the description of the single terms. no nco mm er cia l u se on ly 40 the air2water model: guidelines, challenges, and perspectives seasonal and interannual patterns of thermal stratification, as it has been clearly demonstrated for the cases of lake constance (toffolon et al., 2014a) and lake superior (piccolroaz et al., 2015a). equations (4) and (5) taken together constitute the air2water model in its full, 8-parameter version. two simplified versions of the model are also available: a 6-parameter version where δ=1 when the lake is inversely stratified; and a 4-parameter version which, beyond the above simplification, does not include the externally imposed sinusoidal forcing (i.e., a5=0). this latter version can be considered particularly appropriate when the annual cycles of tw and/or of ta are approximately sinusoidal: in fact, from basic principles of trigonometry, the sum of sinusoidal functions with the same period (i.e., 1 year) but different amplitude and phase, yields another sinusoid with different amplitude and phase but the same period. therefore, two sinusoids are enough, and the term can be removed. for the reason given in the results section (second paragraph), the whole analysis is performed considering only the 4and 6-parameter versions of the air2water model, without loss in generality. numerical solution and model calibration the second release of the air2water model is now available at https://github.com/spiccolroaz/air2water, where the source code (written in fortran 90/95), the precompiled executable files (linux/windows), a readme file, and an example application are freely downloadable (the code is published under the creative commons attribution-sharealike 3.0 license). in this new release, the main improvement concerns the numerical solution of the ordinary differential equation (4), which, together with equations (5), constitutes the air2water model. users can now choose among euler, runge-kutta 2nd order, rungekutta 4th order, and crank-nicolson numerical schemes. the first three schemes are explicit, and in summer, when δ→0, it may happen that a daily time step is too large to adequately integrate equation (4), possibly generating numerical instabilities. in order to avoid this situation and provide an accurate prediction of tw, an adaptive sub-stepping procedure has been implemented, in which the original integration time step of one day is divided into a number of equal sub-steps according to the stability conditions of the method (butcher, 2008). predictions of tw are anyway provided at daily time scale. conversely, the last numerical scheme is implicit, 2nd order accurate, and unconditionally stable: a sub-stepping procedure is not required and the daily time step is used for the whole simulation, making it generally faster (but less accurate than runge-kutta 4th order) than the previous schemes. in this case, in order to obtain a closed-form analytical expression of equation (4), is handled explicitly, thus only to the numerator of the right-hand side of equation (4) has been discretized according to the crank-nicolson scheme. model calibration is performed through a monte carlo-based optimization approach in which a large number of parameter sets are sampled and evaluated in terms of a given metric of model efficiency. here, the root mean square error (rmse) between observed and modelled values is considered as an optimization metric, meaning that at the end of the optimization loop the best set of parameters is identified as the one providing the smallest rmse. the sampling procedure is performed through the particle swarm optimization (pso) algorithm, a simple and powerful population-based stochastic optimization technique firstly proposed by kennedy and eberhart (1995) for solving engineering problems, and successively applied to a variety of different fields, including hydrology (gill et al., 2006; piccolroaz et al., 2015b). for further details about this optimization procedure, the reader is referred to supplementary material b. numerical integration of equation (4) requires that the series of air temperature (i.e., the external forcing) be continuous and at daily resolution. therefore gaps (in case they exist) must be reconstructed e.g., by replacement with the average value of all air temperature measurements available in the data set for the same specific day of the year when the data is missing. conversely, the time series of observed lst can contain missing data. in this case, missing data are not replaced, and they simply do not contribute to the evaluation of the prediction performance (e.g., through the evaluation of rmse between observed and simulated lst). this allows for using air2water with lst observational time series at any frequency (e.g., weekly, monthly, seasonal) that is not necessarily the daily, or simply with irregular time series. the effect on model performance of the presence of missing lst data will be analysed in detail in the results section. as a final note, besides rmse the user can choose between other metrics of model performance: the nash-sutcliffe efficiency index (nse, nash and sutcliffe, 1970) and the kling-gupta efficiency index (kge, gupta et al., 2009). in addition, model calibration can be performed using simple random sampling or the latin hypercube sampling technique (mckay et al., 1979) besides the pso, which are computationally more expensive but explore more uniformly the space of parameters, allowing for conducting sensitivity analyses of model parameters. results evaluating the a priori range of model parameters as mentioned in the methods section, to ensure proper model calibration, model parameters are required to be no nco mm er cia l u se on ly 41s. piccolroaz defined within a physically consistent a priori range of variation. this range should be sufficiently wide to allow for the existence of an optimal and physically plausible set of parameters, and at the same time it should not be indiscreetly large to avoid convergence to unrealistic solutions. suitable a priori ranges of variations for parameters ai,i=1, 2, 3, 5 can be evaluated on the basis of physical considerations, recalling that ai=âi/(dr ρcp). reliable estimates of âi and dr are therefore required. the possible range of variation of parameters âi can be obtained from equations (a11)-(a15) in supplementary material a, considering all possible values and combinations of the physical coefficients that appear in these equations (martin and mccutcheon, 1998). also the reference depth dr , i.e., the mean depth of the largest water volume involved in the surface heat budget of the lake during the year, see methods) can be assumed to vary within a range of possible values. reasonably, dr is bounded from above by the average depth of the lake (d=v/a, where v and a are volume and surface area of the lake, respectively), i.e., when dr=d the whole lake participates to the heat exchange with the atmosphere when the water column is well mixed. however, for the case of very shallow lakes (e.g., having the mean depth on the order of a few meters), fig. 3. estimate of the a priori range of variation of model parameters as a function of the mean depth of the lake , and regression relationships as determined by toffolon et al. (2014a) analyzing 14 lakes with different morphologies. no nco mm er cia l u se on ly 42 the air2water model: guidelines, challenges, and perspectives the effective volume participating to the heat budget may partially involve lake sediments making the effective volume larger than the mere lake water volume (toffolon et al., 2014a). this possibility is implicitly accounted for in the calibration of model parameters without the need of specifying any additional input information, but simply setting the upper bound of dr to be larger than d (10 m is a reasonable and safe choice). as for the lower bound of dr, experience suggests that a simple option is to linearly vary it from d=1 m for m to 50 m for m, which is certainly a conservative underestimate. in fact, in lake baikal (russia, the world’s deepest lake) d=744 m and 50 m only roughly represents the thickness of the epilimnion during strong thermal stratification (piccolroaz and toffolon, 2013), suggesting that the dr is certainly larger than this value. parameter , which is the phase of the sinusoidal term with amplitude a5 summing up all contributions to the heat budget with the exception of the direct effect of air temperature, simply varies from 0 to 1. parameter a4 controls the intensity of the stratification (thus the volume that is affected by the heat exchange), and, based on practical experience, its possible range of variation can be defined as in fig. 3d. fig. 3 shows the range of variation of all parameters as a function of d, evaluated based on the above considerations and setting the coefficients in equations (a11)(a15) according to typical values that they assume in the temperate region. note that in principle this estimate is coherent with the 6and 8-parameter versions of the model, while in the 4-parameter version the meaning of the parameters is slightly different as parameter a5 is absorbed into parameters a1, a2, and a3. in general, however, experience suggests that the same range of parameters can be safely used for all versions of the model. fig. 3 also shows the relationships between model parameters and lake average depth d as determined by toffolon et al. (2014a) where 14 temperate lakes were analysed which were characterized by significantly different morphologies, using the 4-and 8-parameter versions of the model (here the relationships obtained for the full 8-parameter version are assumed valid also for the 6-parameter version given the strong similarity between the two versions of the models). the regressions between model parameters and are d in tab. 2. from the combined analysis of fig. 3 and tab. 2, two main comments can be made: first, the regression lines are well within the physical a priori ranges of parameters, suggesting that these ranges are properly defined. the only exception is parameter a1 in the 6-parameter version, whose regression line is beneath the lower physical bound for d>300 m. however, for such deep lakes, previous results suggest that this relationship is likely not significant (see e.g., the case of lake baikal in the original paper by toffolon et al., 2014a), and in any case the overall dependence on d is weak. second, and perhaps more important, despite by definition parameters ai,i=1, 2, 3, 5, should depend inversely on depth, the regression lines do not simply scale with d–1 (see e.g., the exponents of the power laws in tab. 1). this is indicative that air2water is able to suitably reproduce the complex thermal behaviour of a lake, by transferring the information contained in the observed data directly to model parameters, which, in turn, have a significant dependence on lake depth. post-calibration analysis the optimal set of parameters resulting from the calibration procedure is required to be well centred within the a priori range of variation, in order to exclude any confinement effect due to bounds that are too narrow. this is expected to always be the case when using the a priori range of parameters discussed in the previous section. however, it is always preferable to perform an a posteriori sensitivity analysis, aimed at excluding the eventuality of parameter ranges that are too narrow and at the same time evaluating parameters’ identifiability and significance. this analysis is easily done producing and analysing the shape of the socalled dotty plot”, which are projections of the measure of model performance (in this case expressed through rmse) obtained after the calibration procedure within the hyperspace of parameters, onto single parameter axes (beven and freer, 2001; see fig. 4 for a schematic). preferably, dotty plots should be obtained using simple random sampling or latin hypercube sampling techniques for model calibration instead of pso, to avoid clustering around the best solution. if a dotty plot is sharp and well defined (as in fig. 4a) it means that the parameter is significant and well identifiable, while if it is flat and scattered (as in fig. 4b) it means that the parameter is not significant or the model is overparameterized. detailed discussions about parameters identifiability of the three versions of the air2water model can be found in piccolroaz et al. (2013) and toffolon et al. (2014a). parameters are well identifiable for all versions of the model (being slightly higher in the 4-parameter version tab. 2. equations of the regression relationships between model parameters and the mean depth of the lake found by toffolon et al. (2014a) analysing 14 lakes with different morphologies, and shown in fig. 3. parameter regression equation 4 parameters 6(8) parameters a1 –0.042+0.017 log (d) 0.488–0.096 log (d) a2 0.223 d–0.635 0.207 d–0.672 a3 0.175 d–0.540 0.262 d–0.659 a4 35.4 d–0.360 31.3 d–0.330 a5 – 0.843 d–0.732 a6 – 0.628–0.030 log (d) no nco mm er cia l u se on ly 43s. piccolroaz due to lower number of parameters), with the only exception of parameters a7 and a8 in the full, 8-parameter version. the main reason is that these parameters are not fully independent, and may produce significant interactions. a more appropriate parameterization of δ during inverse stratification and ice formation periods is currently under development. since a7 generally achieves relatively high values implying δ~1 for tw≤4°c (toffolon et al., 2014a), the following analysis is performed considering only the 4and 6-parameter versions, still retaining full generality. results of the 4and 6-parameter versions of the model for the cases of lakes superior and erie are presented in fig. 5 and fig. 6. in both cases, the calibration of the parameters was performed using two-thirds of the data set (13 years, from 1995 to 2007) and leaving one-third for the validation (7 years, from 2008 to 2014). fig. 5 shows scatterplots for the two lakes and the two versions of air2water during the calibration period. no systematic deviation (bias) is observed, and the dispersion along the diagonal does not exhibit significant trends. both these characteristics are confirmed by the relatively small values of rmse and values of the coefficient of determination (r2) close to one: rmse=1.00°c and r2=0.97 and rmse=0.93°c and r2=0.97 for lake superior (4and 6-parameter versions), and rmse=0.87°c and r2=0.99 and rmse=0.82°c and r2=0.99 for lake erie (same model versions). in figure 6 simulated lst is compared with observations during the validation period, showing close agreement overall. rmses in validation are: 0.90°c and 0.79°c for lake superior (4and 6-parameter versions), and 0.73°c and 0.68°c for lake erie (same model versions). fig. 6 displays the ability of the model to appropriately capture seasonal dynamics and interannual variability. this suggests that air2water is a valuable tool for long-term predictions of lst, in both deep and shallow lakes. the model shows slightly weaker performance in the case of lake superior due to its more complex thermal behaviour, which is significantly controlled by stratification and thermal inertia (piccolroaz et al., 2015a). furthermore, the relative worsening of the 4-parameter version relative to the 6-parameter version is higher in this case (rmse increases by 14% in validation) than in lake erie (rmse increases by 7%). this suggests that the hypotheses at the basis of the derivation of the simplest, 4-parameter version of the air2water model (see methods) are likely to be more appropriate in the case of shallow lakes, and anyway when air and water temperature annual cycles shows a nearly sinusoidal pattern (see methods and fig. 1). effects of missing data on model performance in this section, the effect on model performance of the presence of missing data in the time series of observed lst is analyzed and discussed. in fact, long-term continuous observations of lst are only rarely available, thus often limiting their practical use. for example, in lakes that freeze, offshore monitoring buoys are generally removed during winter to prevent damage from ice. also lst time series retrieved from satellite imagery, which are generally more continuous during the year, may have gaps during periods of cloudiness. finally, the constant and continuous in-situ monitoring of a lake requires sufficient funding and qualified personnel which are not always available, especially over long-term periods. the performance of the air2water model is evaluated by progressively increasing the number of gaps in the lst series, from 10% to 90%, by increments of 10%. percentages of missing data of 95%, 97%, 99%, and 99.5% are also considered, which roughly correspond to the availability of 18, 11 (monthly), 4 (seasonal), and 2 measurements per year, on average. in order to perform a robust statistical analysis, for each of the considered missing data scenarios an ensemble of 100 series of lst is obtained from the original, continuous series of observations, by randomly excluding the correspondent number of data. then, the model is calibrated on the basis of these artificially deteriorated 13-year series of data (1995 to 2007), and validated on the remaining 7-year period (2008 to 2014). in order to allow for a fair and unbiased comparison among model performance obtained for the different scenarios and for the reference (i.e., continuous time series, no gaps) simulation, the validation period is not modified and the same continuous series shown in fig. 6 is used in all cases. results of the analysis for both the 4and the 6-parameter versions of the model are shown in fig. 7. for each scenario, the rmses obtained for the ensemble of simulations are presented through a box plot, where the circle indicates the median value of the distribution. by comparing the median values with the rmses of the reference simulations (continuous lines), it is possible to conclude that, as a general tendency, no degradation of model performance will occur until a data gap of about 50%-60% fig. 4. schematic of (a) a sharp and well defined dotty plot and (b) a flat and scattered dotty plot. each black dot corresponds to one model simulation (one parameter set) and the red dot represents the optimal parameter set. no nco mm er cia l u se on ly 44 the air2water model: guidelines, challenges, and perspectives for the 6-parameter version, and until a data gap of about 70% for the 4-parameter version. in any case the whole box plot is within 10% of the reference value until a data gap of about 90%-95%. when the percentage of the data gap is larger, model performance diminishes, which occurs faster for the 6-parameter version of the model and for the deepest lake. in fact, when the data gap is significantly large the structure of the 6-parameter version of the model may become too complex (i.e., there are too many parameters) relative to the number of observations, thus running the risk of overfitting (vapnik, 1999). this is more evident in deep lakes, which are characterized by more complex thermal dynamics due to the significant role played by stratification and thermal inertia (piccolroaz et al., 2015a). it is possible to conclude that the 6-parameter version of the model is preferable to the 4-parameter version when the amount of missing data is lower than 95% (i.e., when data are available at about bi-weekly resolution, on average). up to 95% missing data, the model still performs reasonably well compared to the reference case when the lst series in calibration is complete. with more than 95% of data missing, the air2water model should be used cautiously, making a case by case assessment evaluating whether results are reasonable compared to the expected behaviour of the lake, and preferring the simplest 4-parameter version. in particular, this version of the model shows acceptable performance until the percentage of missing data reaches about 97% (i.e., when data are available at about monthly resolution, on average), and particularly for the shallow lake erie. as a final remark, note that some boxplots in fig. 6 are partially (and to a minor extent) beneath the reference value of rmse, which indicates that there are a few cases where the optimal set of parameters obtained with a less complete series of lst observations provide slightly better performances in validation. this is likely due to the specific time period considered in the analysis and to the quality of lst observations, and is not explored further here. how length of the calibration period and percentage of missing data affect model performance the analysis presented in the previous section is specific of a 13-year long calibration period, and here it is generalized by considering different lengths of the calibration period, with the aim to provide an overview of the consequences of data scarcity on model performance. the final aim is to provide the user of the air2water model with a criterion to assess whether the observational dataset used for model calibration is sufficiently informative to obtain a reliable calibration or not. the same analysis described above is therefore extended considering different lengths of the calibration period: 1, 2, 3, 5, 8, and 13 years (as a tribute to leonardo fibonacci). in order not to introduce biases in the results, when testing calibration periods shorter than 13 years, the sequences of years are randomly extracted from the original 13-year long series ranging from 1995 to 2007. then, in analogy with the previous analysis, an ensemble of 100 artificially deteriorated series of lst is randomly generated for each combination of percentage of gaps and length of the calibration period. 6 calfig. 5. scatter plot of observed against simulated lst during the calibration period (1995-2007) for (a) lake superior and (b) lake erie, and for the 4and 6parameter versions of the air2water model. no nco mm er cia l u se on ly 45s. piccolroaz ibration period lengths and 9 percentages of missing data are investigated for a total of 54 different combinations (hereafter referred to as scenarios) and 5400 model runs. results are presented in fig. 8, which shows the relative deterioration of each scenario with respect to the best performing case (through the ratio rmsei/min ({rmsei}54i =1), where rmsei is the median root mean square error of the i-th scenario in validation, and ranges from 1 to 54), for the two lakes and the two versions of the model. results confirm and extend the previous analysis: a larger degradation (in relative terms) of model performance with increasing deterioration of the dataset is observed for the 6-parameter version (and, secondarily, for the deepest lake). in this case, at least 8 years of data with no more than 80% of missing data are required to avoid a worsening of more than 10% from the best scenario, for both lake superior and lake erie. conversely, with the 4-parameter version a calibration period of 2 or 3 years with up to 80% or 90% missing data is sufficient to obtain the same deterioration in model performance (again in relative terms), for lake superior and lake erie, respectively. furthermore, in general, similar model performances can be achieved with a lower number of total observations (i.e., larger percentage of missing data) if the calibration period is longer. in other words, a longer calibration period with fewer measurements may be more informative than a shorter calibration period with more data, suggesting the high value of disposing of a series of data characterized by significant interannual variability. as an example, model performance is roughly the same when fig. 6. comparison between simulated and observed surface water temperature during the validation period (2008–2014) for (a) lake superior and (b) lake erie, and for the 4and 6parameter versions of the air2water model. observed air temperature data are also presented. fig. 7. box plots of rmses values obtained in validation considering different percentages of missing data in the calibration time series of lst, for (a) lake superior and (b) lake erie. the circle indicates the median value of the distributions. for each missing data scenario, an ensemble of 100 artificially deteriorated series of lst is randomly generated. no nco mm er cia l u se on ly 46 the air2water model: guidelines, challenges, and perspectives considering a 13-year long period with 95% of gaps (i.e., 237 valid data) or a 8-year long period with 80% of gaps (i.e., 584 valid data; see fig. 8b). finally, rmsei obtained using the 4-paramters and 6parameters versions of the model are compared for the two lakes, making possible to draw a map of preference (in absolute terms) between the two versions of the model as a function of the different scenarios (see fig. 9). in both cases, the 4-paramters version of the air2water model is more performant, thus it is to be preferred, versus the 6parameter version when the calibration period is shorter than about 5 years, or when it is longer but with more than 97% of gaps. the same considerations about model overfitting discussed in the previous section apply also here. discussion in previous works, piccolroaz et al. (2013, 2015a) and toffolon et al. (2104a) have already demonstrated the high potential of the air2water model as a simple, yet effective, predictive tool for simulating lst when only air temperature data are available. the model is able to properly simulate the hysteresis loop between air and water temperature in both shallow and deep lakes, and to accurately capture seasonal and interannual fluctuations of lst. the model also allows for the simulation of stratification dynamics in lakes, without the need to introduce a complex description of the air-water interface processes fig. 8. air2water model performance (in terms of increasing rmse in validation) as a function of the amount of missing data and calibration period length, for lake superior and lake erie, and for the 4and 6parameter versions. no nco mm er cia l u se on ly 47s. piccolroaz based on a detailed quantification of the single heat flux components. furthermore, it has been successfully applied using different sources of data, as e.g., lst measured at buoys or retrieved from satellite and air temperature from observations or re-analysis, suggesting a high degree of flexibility concerning the possibility to use different types of data as input. this is possible because of the physically-based structure of the model allowing for the acquisition of information about the studied system directly from the data, through the calibration of model parameters. this process is further facilitated given the extreme simplicity of the air2water model, which makes it particularly prone to automatic calibration procedures within a monte carlo-like framework. in this way, model parameters assimilate the information contained in the observations, and in turn the user may learn how the real system behaves from the values of the parameters, identifying what are the most important processes controlling the thermal response of the lake. informativeness of observations is a crucial aspect that should be considered carefully in order to exclude an improper calibration of the model parameters, and an unreliable, or at least uncertain, prediction of lst. this critical detail is addressed in the results section, where air2water model users can find some recommended best practices for a proper use of the model. the simplicity and robustness of the air2water model suggest its possible use in different context and for different purposes, heading towards new challenges: • the investigation of the response of lakes to air temperature variations under climate change scenarios. in this perspective, the air2water model represents a valuable alternative tool to simpler regression models, which require the same data in input but are not able to address some fundamental processes (e.g., the hysteresis cycle between air and water temperature); but also to more complex process-based models, which require a significantly larger amount of input data without showing significantly better performances (see e.g., results in thiery et al., 2014). • the direct coupling with atmospheric circulation and weather forecasting models. recent attempts in this direction have been made adopting complex one-dimensional lake models (e.g., using k-e turbulence model as in goyette and perroud 2012), but have inevitably shown some limitations as e.g., expensive computational cost and the need of ground-truth information. simpler models have also been used to this aim (dating back to hostetler et al., 1993), but in any case requiring the entire set of meteorological data. again, the simplicity, parsimony, and robustness of air2water make it a good candidate for being adopted as a lumped lake model integrated in meteorological models. • the coupling with simple water quality, ecological and biogeochemical modules in order to investigate processes that are significantly controlled by water temperature, as e.g., nutrients, dissolved oxygen, and aquatic ecosystem dynamics. this would be a good opportunity to cross the boundaries (according to toffolon et al., 2014b) between the various disciplines of aquatic science facilitating the dialogue and collaboration between scientists from different background. indeed, fragmentation of limnology into expert, specialised fields, with limited interaction is a wellknown major issue of this branch of science (peters, 1990; lewis, 1995; salmaso and mosello, 2010). • the definition of regionalization relationships between model parameters and morphological characteristics of lakes, with the final aim to apply the model to ungauged lakes. expanding the analysis of toffolon et al. (2014a) that analysed 14 temperate lakes characterized by different morphology, by including additional lakes possibly at different latitudes (e.g., tropical and polar lakes) is particularly interesting. in this regard, the growing availability of collections of lakes’ observational data at the global scale is particularly attractive (e.g., global lake temperature collaboration gltc, sharma et al., 2015; global lake ecological observatory network gleon, weathers et al., 2013), also for testing the air2water model on lakes outside of the temperate zone (e.g., tropical or polar lakes). furthermore, the application of air2water globally may provide interesting insights into how lst in lakes around the world is expected to respond to climate change in the future, possibly identifying some meaningful hotspots as in o’reilly et al. (2015). fig. 9. diagram of preference between the 4and 6parameter versions of the air2water model as a function of the amount of missing data and calibration period length, for lake superior and lake erie. no nco mm er cia l u se on ly 48 the air2water model: guidelines, challenges, and perspectives conclusions the results of this work provide the reader with guidelines and best practices for using the air2water model, as a simple tool to predict lst when only air temperature is available. after having briefly recalled the derivation of the model and the meaning of parameters, the model is used to simulate lst in two lakes characterized by significantly different depths: lake superior and lake erie (usa). these two case studies are chosen as clear and emblematic examples of a deep and a shallow temperate lake characterized by markedly different thermal responses to external forcing, with the aim of making the results of the analysis as much as possible general and comprehensive. the whole analysis is carried out considering the 4and 6-parameter versions of the model. the full, 8-parameter version is not considered here, due to the sub-optimal parameterization of during inverse stratification and ice formation periods, whose improvement is currently under development. in this work, the possible user of the air2water model is provided with all the fundamental information for a proper use of the model: from the initial definition of appropriate a priori range of variations of model parameters to an effective post-processing analysis of results, passing through a sensitivity analysis about the influence of missing data on model performance. particular attention is paid to this last point, which can be summarized as follows: i) longer calibration periods with overall less number of measurements is likely to be more informative than shorter calibration periods with more data (suggesting the high value of disposing of time series with high interannual variability); ii) when the number of missing data increases, model performance diminishes more for the 6-parameter version, suggesting the risk of model overfitting; iii) for short calibration time series (e.g., shorter than about 5 years in this case), the 4-parameter version of the model is likely to be preferable anyway; and iv) as a secondary effect, model performance diminishes more for deeper lakes when data are missing, compared to shallow lakes, due to complex thermal behaviour that is chiefly influenced by lake depth. coherently with one of the main goals of this work, which is to foster the dialogue among the several branches of aquatic science, a flowchart of the main modelling steps is shown in fig. 10, which is intended to make the sequence of the operational phases at the basis of the use of air2water clearer and easier to follow also to users with different mathematical and/or technical backgrounds. indeed, the air2water model has been developed with the clear intention to offer a simple tool that can indifferently be used by physicists and biologists, modellers and experimentalists, possibly generating new collaborations towards an integrated understanding of how lst responds to climate forcing and what are the effects on the ecological status of the lake. in this perspective, everyone that is interested can collaborate to improve the model with comments, suggestions and contributions, which are highly welcomed and easy to share through https://github.com/spiccolroaz/air2water. acknowledgments the author is grateful to marco toffolon for discussions on an earlier version of the manuscript, to elisa calamita for preliminary analysis of the data, and to ulrike obertegger (edmund mach foundation, italy) for rewriting the post-processing script in r (available on https://github. com/spiccolroaz/air2water). the author is also thankful to noaa (national oceanic and atmospheric administration) for lst data used in this work (data can be downloaded from http://www.glerl.noaa.gov/glsea/asc_1024/) fig. 10. flowchart of the main modelling steps: input data, definition of the a priori ranges of model parameters, run of air2water within a monte carlo optimization framework, results. for a more detailed description of how to use the model, please refer to the file readme.txt in https://github.com/spiccolroaz/air2water. no nco mm er cia l u se on ly 49s. piccolroaz and to ecmwf (european centre for medium-range weather forecasts) for daily air temperature (data can be downloaded from http://apps.ecmwf.int/datasets/data/interim-full-daily/levtype=sfc/). finally, the author thanks the two anonymous reviewers for their constructive comments, which helped to improve the manuscript. references adrian r, o’reilly cm, zagarese h, baines sb, hessen do, keller w, livingstone dm, sommaruga r, straile, d, van donk e, weyhenmeyer ga, winder m, 2009. lakes as sentinels of climate change. limnol. oceanogr. 54:2283-2297. beven k, freer j, 2001. a dynamic topmodel. hydrol. process. 15:1992-2011 butcher jc, 2008. numerical methods for ordinary differential equations. 2nd ed. j. wiley & sons, london. butcher jb, nover d, johnson te, clark cm, 2015. sensitivity of lake thermal and mixing dynamics to climate change. climatic change. 129:295-305. de senerpont domis ln, elser jj, gsell as, huszar vlm, ibelings bw, jeppesen e, kosten s, mooij wm, roland f, sommer u, van donk e, winder m, lürling m, 2013. plankton dynamics under different climatic conditions in space and time. freshwater biol. 58:463-482. flaim g, obertegger u, anesi a, guella g, 2014. temperatureinduced changes in lipid biomarkers and mycosporine-like amino acids in the psychrophilic dinoflagellate peridinium aciculiferum. freshwater biol. 59:985-997. gallina n, salmaso n, morabito g, beniston m, 2013. phytoplankton configuration in six deep lakes in the peri-alpine region: are the key drivers related to eutrophication and climate? aquat. ecol. 47:177-193. gill mk, kaheil yh, khalil a, mckee m, bastidas l, 2006. multiobjective particle swarm optimization for parameter estimation in hydrology. water resour. res. 42:w07417. goyette s, perroud m, 2012. interfacing a onedimensional lake model with a single-column atmospheric model: application to the deep lake geneva, switzerland. water resour. res. 48:w04507. gupta hv, kling h, yilmaz kk, martinez gf, 2009. decomposition of the mean squared error and nse performance criteria: implications for improving hydrological modelling. j. hydrol. 377:80-91. hostetler sw, bates gt, giorgi f, 1993. interactive coupling of a lake thermal model with a regional climate model. j. geophys. res. 98:5045-5057. kennedy j, eberhart r, 1995. particle swarm optimization, p. 1942-1948. proc. ieee int. conf. on neural networks, university of western australia, perth, australia. kettle h, anderson rtnj, livingstone dm, 2004. empirical modeling of summer lake surface temperatures in southwest greenland. limnol. oceanogr. 49:271-282. lewis wm, 1995. limnology, as seen by limnologists. j. contemp. water res. educ. 98:4-8. livingstone dm, lotter af, 1998. the relationship between air and water temperatures in lakes of the swiss plateau: a case study with palaeolimnological implications. j. paleolimnol. 19:181-198. livingstone dm, padisák j, 2007. large-scale coherence in the response of lake surface-water temperatures to synopticscale climate forcing during summer. limnol. oceanogr. 52:896-902. martin jl, mccutcheon s, 1998. hydrodynamics and transport for water quality modeling. crc press. martynov a, sushama l, laprise r, 2010. simulation of temperate freezing lakes by one-dimensional lake models: performance assessment for interactive coupling with regional climate models. boreal environ. res. 15:143-164. mccombie am, 1959. some relations between air temperatures and the surface water temperatures of lakes. limnol. oceanogr. 4:252-258. mckay md, beckman rj, conover wj, 1979. a comparison of three methods for selecting values of input variables in the analysis of output from a computer code. technometrics. 21:239-245. nash je, sutcliffe jv, 1970. river flow forecasting through conceptual models 1. a discussion of principles. j. hydrol. 10: 282-290. o’reilly cm, sharma s, gray dk, hampton se, read js, rowley rj, schneider p, lenters jd, mcintyre pb, kraemer bm, weyhenmeyer ga, straile, d, dong b, adrian r, allan mg, anneville o, arvola l, austin j, bailey jl, baron js, brookes jd, de eyto e, dokulil mt, hamilton dp, havens k, hetherington al, higgins sn, hook s, izmest’eva lr, joehnk kd, kangur k, kasprzak p, kumagai m, kuusisto e, leshkevich g, livingstone dm, macintyre s, may l, melack jm, mueller-navarra dc, naumenko m, noges p, noges t, north rp, plisnier pd, rigosi a, rimmer a, rogora m, rudstam lg, rusak ja, salmaso n, samal nr, schindler de, schladow sg, schmid m, schmidt sr, silow e, soylu me, teubner k, verburg p, voutilainen a,watkinson a, williamson ce, zhang g, 2015. rapid and highly variable warming of lake surface waters around the globe. geophys. res. lett. 42:773-781. perroud ms, goyette s, martynov a, beniston m, anneville o, 2009. simulation of multiannual thermal profiles in deep lake geneva: a comparison of one-dimensional lake models. limnol. oceanogr. 54:1574-1594. peters rh, 1990. pathologies in limnology. mem. ist. ital. idrobiol. 47:181-217. piccolroaz s, toffolon m, 2013. deep water renewal in lake baikal: a model for long term analyses. j. geophys. res.oceans 118:6717-6733. piccolroaz s, toffolon m, majone b, 2013. a simple lumped model to convert air temperature into surface water temperature in lakes. hydrol. earth syst. sci. 17:3323-3338. piccolroaz s, toffolon m, majone b, 2015a. the role of stratification on lakes’ thermal response: the case of lake superior. water resour. res. 51:7270-7288. piccolroaz s, majone b, palmieri f, cassiani g, bellin a, 2015b. on the use of spatially distributed, time-lapse microgravity surveys to inform hydrological modeling, water resour. res. 51:7878-7894. salmaso n, mosello r, 2010. limnological research in the deep southern subalpine lakes: synthesis, directions and perspectives. adv. oceanogr. limnol. 1:29-66. schabhüttl s, hingsamer p, weigelhofer g, hein t, weigert a, striebel m, 2013. temperature and species richness effects no nco mm er cia l u se on ly 50 the air2water model: guidelines, challenges, and perspectives in phytoplankton communities. oecologia 171:527-536. sharma s, walker sc, jackson da, 2008. empirical modelling of lake water-temperature relationships: a comparison of approaches. freshwater biol. 53:897-911. sharma s, gray dk, read js, o’reilly cm, schneider p, qudrat a, gries c, stefanoff s, hampton se, hook s, lenters jd, livingstone dm, mcintyre pb, adrian r, allan mg, anneville o, arvola l, austin j, bailey j, baron js, brookes j, chen y, daly r, dokulil m, dong b, ewing k, de eyto e, hamilton dp, havens k, haydon s, hetzenauer h, heneberry j, hetherington al, higgins sn, hixson e, izmest’eva lr, jones bm, kangur k, kasprzak p, köster o, kraemer bm, kumagai m, kuusisto e, leshkevich g, may l, macintyre s, müller-navarra d, naumenko m, noges p, noges t, niederhauser p, north rp, paterson am, plisnier pd, rigosi a, rimmer a, rogora m, rudstam l, rusak ja, salmaso n, samal nr, schindler de, schladow g, schmidt sr, schultz t, silow ea, straile d, teubner k, verburg p, voutilainen a, watkinson a, weyhenmeyer ga, williamson ce, woo kh, 2015. a global database of lake surface temperatures collected by in situ and satellite methods from 1985-2009. sci. data 2:150008. thiery w, stepanenko vm, fang x, jöhnk kd, li z, martynov a, perroud m, subin, zm, darchambeau f, mironov d, van lipzig npm., 2014. lakemip kivu: evaluating the representation of a large, deep tropical lake by a set of 1-dimensional lake models. tellus ser. a 66:21390. toffolon m, piccolroaz s, majone b, soja am, peeters f, schmid m, wüest a, 2014a. prediction of surface water temperature from air temperature in lakes with different morphology, limnol. oceanogr. 59:2185-2202. toffolon m, piccolroaz s, bouffard d, 2014b. crossing the boundaries of physical limnology. eos 95:403. vapnik vn, 1999. an overview of statistical learning theory. ieee t. neural network 10:988-999. weathers kc, hanson pc, arzberger p, brentrup j, brookes jd, carey cc, gaiser e, hamilton dp, hong gs, ibelings bw, istvánovics v, jennings e, kim b, kratz tk, lin f-p, muraoka k, o’reilly c, piccolo mc, rose kc, ryder e, zhu g, 2013. the global lake ecological observatory network (gleon): the evolution of grassroots network science. bull. limnol. oceanogr. 22:71-73. webb ms, 1974. surface temperatures of lake erie. water resour. res. 10:199-210. wetzel rg, 2001. limnology: lake and river ecosystems. 3rd ed. academic press. winder m, sommer u, 2012. phytoplankton response to a changing climate. hydrobiologia 698:5-16. no nco mm er cia l u se on ly layout 1 dr. alvise vittori was a well-known italian ichthyologist, limnologist and all-round ecologist. he started his professional career as a consultant for commercial fisheries and in 1967 he was hired as an aquatic biologist for the trentino south tyrol region and transferred to the agricultural and forestry experiment station at san michele all’adige, trento in 1972. after many years of service, he was appointed director of the experiment station, where he oversaw its transition to the now fondazione edmund mach. while trained in forestry science (university of florence), alvise’s heart was aquatic, a state of mind that he attributed to growing up on the shores of lake garda. his professional interests ranged from all aspects of fisheries to lake and stream monitoring to habitat restoration. in our small province of trento (italy), home to over 350 lakes, >5000 km of running waters and >60 commercial fish farms, he was a well-known figure, quite active in community affairs. alvise vittori was nothing if not pragmatic and his expert advice on waterrelated issues was often sought out by local administrators and fishermen. we remember most his broad geographical and historical knowledge of the trentino, his ‘hands on’ approach to limnology, his directness, his dry humor and his intolerance of bureaucracy (and bureaucrats). alvise was definitely someone who cannot be forgotten. this was a man with a vision! all who had the opportunity to know alvise vittori can attest to his foresight and intuition. several of his projects, very innovative at the time, are common practices in italy today. a few examples: he was instrumental in starting lake monitoring in trentino (1973) and in the 1980s he founded the san nicolò laboratory on the shores of lake garda where the monitoring program is still continuing. alvise was the force behind several trentino lake restoration projects, most notably lake caldonazzo. as an ecologist, alvise understood the importance of biological monitoring not only of lakes but also of streams and rivers. he initiated the ecological mapping of trentino’s waterways using macroinvertebrates in the early 1980s. he was also instrumental in organizing courses in the use of macroinvertebrates as monitoring tools and promoted the publication of taxonomic guides for alpine and italian macroinvertebrates. here, he worked hand in hand with cisba (centro italiano studi di biologia ambientale) to promote the use of biological indicators as monitoring tools at the national level. alvise’s special rapport with fish and fishermen culminated in the first fisheries management plan (carta ittica -1983) in italy. this pioneering carta ittica was soon copied by other provinces in italy. he was acutely aware of the importance of healthy fish stock both for commercial fish farms and for restocking programs in rivers and streams. with this in mind he founded the experimental fish farm at the fondazione edmund mach. his great contribution to inland fisheries management was acknowledged by being named honorary member of the italian freshwater icthyologist society (aiiad associazione italiana ittiologi acque dolci). alvise’s legacy lives on, and in many ways, he helped define applied limnology in italy. many of his initiatives are still ongoing, either here, at the fondazione e. mach, or at the trento environmental protection agency (appa). alvise vittori was a mentor not only to us, but to dozens of trentino students and fledgling limnologists. more than a scientist, he saw himself as a public servant and was always guided by the need to explain his science and the spending of public monies to ordinary citizens. he shunned formal scientific journals, with most of his scientific production published in esperienze e ricerche, the yearly publication of the former agricultural and forestry experiment station (now fondazione edmund mach), but was a keen and down to earth communicator to the general public. in fact, almost all of alvise’s writings were published in local journals and trade magazines. alvise retired from public service in 1997 and dedicated his retirement to teaching ecology to senior citizens, gardening and truffle hunting. he is survived by his wife elvira, his two children and his beloved grandchildren. selected publications bovo g, vittori a, barrucchelli g, 1986. [carriers of viral haemorrhagic septicemia and infectious pancreatic remembrance remembering alvise vittori (riva del garda 1932 trento 2018) no nco mm er cia l u se on ly g. flaim and f. ciutti16 necrosis viruses among salmonids in the sarca river from pinzolo to the ponte pià basin near salmon farms].[article in italian]. esp. ric. 15: 229-236. chemini c, vittori a, 1976. [a case of a massive appearance of collembola in trentino].[article in italian]. nat. alp. 27:46-51. flaim g, leonardi g, pinamonti v, sittoni l, vittori a. 1995.[ mapping the biological quality of running waters in the province of trento].[article in italian]. quaderni esp. ric. 1:1-125. flaim g, pinamonti v, vittori a, 1980. [the acute and semi-chronic toxicity of a mixture of pesticides commonly used in fruit orchards for the fish carassius carassius l.].[article in italian]. esp. ric. 9:135-146. maiolini b, vittori a, 1980. [on the presence of holopedium gibberum zaddack in lake valbona, province of trento: (crustacea: cladocera)].[article in italian]. st. trent. sci. nat. acta biol. 57:75-81. pat (autonomous province of trento), 1983. [the fish management plan: the foundation document of the autonomous province of trento for a rational water management policy].[book in italian]. provincia autonoma di trento. editrice temi, trento: 117 pp. pontalti l, vittori a, 1996. [pilot plant know-how in relation to salmonid production, p. 295-305].[article in italian]. proceedings vi national congress aiiad. pontalti l, vittori a, 1999. [improving stream quality in trentino by creating artificial lagoons, p. 46-58]. proceedings vii national congress aiiad. pontalti l, vittori a, 1999. [improving the biological quality of a stream in trentino by creating an artificial lagoon].[article in italian]. dendronatura 19:37-44. pontalti l, vittori a, 2001. [growth of the garda carpione (salmo carpio l.) in aquaculture, p. 97-102].[article in italian]. proceedings viii national congress aiiad. vittori a, 1966. [salvelinus alpinus sch. and salmo marmoratus cuv].[article in italian]. nat. alp. 17:39-44. vittori a, 1968. [current economy of salmonid farming in the region trentino south tyrol].[article in italian]. econ trent 1:1-7. vittori a, 1969. [ecological surveys relating to alterations of the lacustrine biocoenosis of lake tovel (trentino)] [article in italian]. st. trent. sci. nat. 46b:267-281. vittori a, 1972. [biological problems related to the lack of red blooms in lake tovel].[article in italian]. nat. alp. 23:108-113. vittori a, 1973. [hydrobiochemical survey of lake tovel during the 1971-1972 sampling campaign].[article in italian]. esp. ric. 3:233-246. vittori a, 1973. [manual for the first survey of pollution of running waters in trentino-alto adige according to d.p.g.r. june 15, 1970, n. 1483].[book in italian]. stazione sperimentale agraria forestale, san michele all’adige. vittori a, 1974. [limnological observations on the principal lakes in the province of trento].[article in italian]. esp. ric. 3:433-449. vittori a, 1975. [one of the ecological factors that governs salmonid growth in fish farms: dissolved oxygen].[article in italian]. econ. trent. vittori a, 1977. [limnological observations on the principal lakes in the province of trento].[article in italian]. esp. ric. 6:389-471. vittori a, 1978. [the italian experience: lake caldonazzo].[article in italian]. press conference on limnology, circolo della stampa, milano. vittori a, 1981. [considerations on the hydrobiology of the trentino part of monte baldo].[article in italian]. nat. alp. 32:47-49. vittori a, 1981. [multi-year experimentation on artificial fertilization and incubation of native salmonids].[article in italian]. esp. ric. 10:193-200. vittori a, 1982. [the fish management plan is a basic tool for understanding water productivity].[article in italian]. qui touring corriere della sera 12:57-58. vittori a, 1985. [lake trout (salmo lacustris l), the queen of the benaco].[article in italian]. nat. alp. 36:43-44. vittori a, 1989. [limnological observations and the trophic state of six lakes in trentino: study period 1988-1989].[article in italian]. esp. ric. 19:295-325. vittori a, 1992. [fish: the biology, morphology and distribution of fish species that inhabit trentino waters].[book in italian]. provincia autonoma di trento: 88 pp. vittori a, leonardi g, pinamonti v, siligardi m, 1987. [lago di caldonazzo: further limnological observations for the years 1985-1986].[article in italian]. esp. ric. 16:149-192. vittori a, leonardi g, pinamonti v, siligardi m, 1992. [lake caldonazzo: limnological observations for the years 1987-1988].[article in italian]. esp. ric. 18:199-227. vittori a, siligardi m, ciutti f, 1999. [training in the use of stream macroinvertebrates as a monitoring tool (i.b.e): what we have learned after 13 years of courses, p. 237-238].in: g.n. baldaccini and g. sansoni (eds.), [biologists and the environment: towards the new millennium].[book in italian]. reggio emilia: cisba. vittori a, siligardi m, leonardi g, pinamonti v, 1983. [a dedicated abacus as a tool for assessing the quality of running water].[article in italian]. esp. ric. 12: 155-163. vittori a, siligardi m, leonardi g, pinamonti v, 1985. [limnological observations and the trophic state of eight lakes in trentino].[article in italian]. esp. ric. 14: 155-163. no nco mm er cia l u se on ly remembering alvise vittori 17 vittori a, siligardi m, pinamonti v, flaim g, leonardi g, 1985. [lake caldonazzo: an experience of applied limnology from 1975 to 1984].[book in italian]. stazione sperimentale agraria forestale, san michele all’adige: 255 pp. vittori a, siligardi m, leonardi g, pinamonti v, 1987. [brief note on the fish population in lake caldonazzo].[article in italian]. esp. ric. 16:143-148. giovanna flaim, francesca ciutti fondazione edmund mach san michele all’adige (tn), italy corresponding author: giovanna.flaim@fmach.it received: 15 may 2018. accepted: 23 may 2019. this work is licensed under a creative commons attribution noncommercial 4.0 license (cc by-nc 4.0). ©copyright: the author(s), 2019 licensee pagepress, italy advances in oceanography and limnology, 2019; 10:8286 doi: 10.4081/aiol.2019.8286 no nco mm er cia l u se on ly layout 1 introduction freshwater ecosystems are tightly linked to the surrounding terrestrial landscape, which continuously exports soil material and microorganisms to the water. understanding the functioning and structuring of lake and river bacterioplankton communities therefore requires taking into consideration these potential linkages between habitats. however, most current work on microbial biogeography is restricted to single types of ecosystems (i.e., only lakes, only soils, only rivers), thus disregarding the potential dispersal of taxa between communities from similar or different habitats, and rendering a rather fragmented view of microbial biogeography. the few recent studies considering such land-water linkages have shown not only that this connectivity with the terrestrial environment may influence aquatic microbial communities by providing nutrients and dissolved organic matter of different quality (besemer et al., 2013; berggren and del giorgio, 2015; ruiz-gonzález et al., 2015a; wilhelm et al., 2015), but also that the immigration of microbes from the surrounding catchment can largely explain the local composition of the receiving community. for example, terrestrial runoff causes the advection of high numbers of bacterial taxa to aquatic bodies, and consequently the systems most tightly connected to the landscape will be more strongly influenced by this transport of terrestrial bacteria. although headwater streams have been suggested to act as integrators and vectors of microbial diversity from soil sources (crump et al., 2007; nelson et al., 2009; besemer et al., 2012, 2013), only a few studies have directly explored this (savio et al 2015; ruiz-gonzález et al., 2015a; niño-garcía et al 2016a), showing that most of these dispersed bacteria, which dominate communities in headwaters, gradually disappear along the river continuum. this results in a diadvances in oceanography and limnology, 2017; 8(2): 222-234 article doi: 10.4081/aiol.2017.7232 this work is licensed under a creative commons attribution-noncommercial 4.0 international license (cc by-nc 4.0). contrasting dynamics and environmental controls of dispersed bacteria along a hydrologic gradient clara ruiz-gonzález,1* juan pablo niño-garcía,2 martin berggren,3 paul a. del giorgio4 1institut de ciències del mar (icm-csic), passeig marítim de la barceloneta 37-49, 08003, barcelona, spain; 2escuela de microbiología, universidad de antioquia, ciudad universitaria calle 67 no 53-108, medellín, colombia; 3department of physical geography and ecosystem science, lund university, sölvegatan 12, se-223 62, lund, sweden; 4groupe de recherche interuniversitaire en limnologie et en environnement aquatique (gril), département des sciences biologiques, université du québec à montréal, case postale 8888, succursale centre-ville, montréal, qc, h3c 3p8, canada *corresponding author: clara.ruiz.glez@gmail.com abstract freshwater bacterioplankton communities are influenced by the transport of bacteria from the surrounding terrestrial environments. it has been shown that, although most of these dispersed bacteria gradually disappear along the hydrologic continuum, some can thrive in aquatic systems and become dominant, leading to a gradual succession of communities. here we aimed at exploring the environmental factors driving the structure of such contrasting bacterial populations as well as their functional properties. using illumina sequencing of the 16s rrna gene, we characterized the taxonomic composition of bacterioplankton communities from 10 streams and rivers in québec spanning the whole hydrologic continuum (river strahler order 0 to 7), which were sampled in two occasions. with the aim to understand the fate and controls of the transported bacteria, among the taxa present at the origin of the hydrologic gradient (i.e., in the smallest headwater streams) we identified two types of dynamics: i) ‘tourist’ taxa, which were those that decreased in abundance from the headwaters towards the largest rivers, and ii) ‘seed’ taxa, those that increased their abundances along the hydrologic continuum. communities changed gradually from the fast-flowing headwater streams dominated by ‘tourist’ taxa (ca. 95% of the sequences) towards the largest rivers (strahler order 4-7) where ‘seed’ taxa comprised up to 80% of community sequences. variation in taxonomic composition of the communities dominated by ‘tourist’ taxa in streams seemed related to different degree of terrestrial inputs, whereas compositional changes in ‘seed’ communities in the large rivers were linked to differences in autochthonous processes. finally, the two types of communities differed significantly in their metabolic potential assessed through biolog ecoplates. all this suggests that hydrologic transport modulates the gradual replacement of two contrasting population types subjected to different environmental controls and with different metabolic potentials. moreover, we show that the separate exploration of the two pools of taxa allows unveiling environmental drivers and processes operating on them that remain hidden if explored at the whole community level. key words: dispersal; river continuum; environmental sorting; aquatic bacterial communities; illumina sequencing. received: december 2017. accepted: december 2017. no nco mm er cia l u se on ly 223c. ruiz-gonzález et al. rectional structuring of aquatic communities and a gradual decrease in taxonomic richness that has been observed both along individual rivers (savio et al 2015) or across multiple unconnected aquatic sites (ruiz-gonzález et al., 2015a; niño-garcía et al 2016a). more importantly, however, some of these transported soil bacteria seem to be able to grow during transit in the water and to dominate freshwater bacterioplankton communities when the water residence time is long enough, i.e., in large rivers and lakes (crump et al 2012, ruiz-gonzález et al., 2015a). the residence time and the distance travelled by a given parcel of water will thus determine the relative proportions of these two contrasting pools of communities that move along the aquatic continuum. the most basic distinction is between the ones that thrive and adapt to local conditions (i.e., ‘seed’ taxa, sensu ruiz-gonzález et al., 2015a) and those that are mal-adapted to the aquatic environment and which therefore decrease in abundance along the hydrologic gradient due to death or dilution (i.e., ‘tourist’ taxa, sensu newton et al., 2011). in support of this idea, several studies have highlighted the role of hydrology, and water residence time in particular, as a major driver of the assembly of aquatic communities in freshwater systems. for example, it was shown that the difference between bacterioplankton communities in lakes and the immigrant assemblages arriving through their inlets varies as a function of lake water residence time (lindström and bergström 2004; lindström et al., 2006). in this regard, crump et al. (2007) suggested that bacteria transported by advection from upstream sources should constitute a small fraction of the total cells in systems with water residence times longer than bacterial doubling times. other studies have explicitly explored the dynamics of different groups of bacterial taxa, showing that, whereas communities at the origin of the hydrologic continuum (i.e., in the smallest headwater streams) are mostly dominated by taxa washed from soils that likely cannot thrive in aquatic systems, downstream systems with longer water residence times mostly comprise typical freshwater taxa adapted to local conditions (savio et al., 2015; ruiz-gonzález et al., 2015a; niño-garcía et al., 2016a). related to this, niño-garcía et al., (2016a) found that, although ph seemed to shape bacterial communities along the whole aquatic continuum in the boreal landscape (ca. 400 streams, rivers and lakes), its role as a driver of taxonomic differences was much more important in large rivers and lakes than in headwater streams where water residence times are shorter. the observed response to ph in fast-flowing streams was interpreted as a signature of the landscape past environmental conditions (niño-garcía et al., 2016a) rather than to in-stream selection. since this was done considering the overall communities, which harbour both types of populations, it is possible that more clear links between environmental drivers and taxonomy will appear considering the two populations separately. for example, catchment properties should be more important in shaping the observed taxonomic composition of ‘tourist’ taxa in headwater streams communities (lear et al., 2013), whereas communities from large rivers or lakes should be mostly structured by local aquatic conditions (logares et al., 2013; souffreau et al., 2015). the corollary of the above is that the succession and replacement of populations with increasing water residence time should be accompanied by shifts in the functional capacities of the communities, since the selection and growth of ‘seed’ taxa along the aquatic continuum should be mostly driven by environmental factors that require a specific functional response. this link between the functional and taxonomic successions that occur along the aquatic continuum may be blurred by the presence of a large fraction of ‘tourist’ taxa that do not necessarily contribute to community functioning. in this regard, niñogarcía et al. (2016b) showed that the presence of most rare taxa across lake bacterioplankton assemblages was due to hydrologic transport from the associated rivers and, although they may not be functional, they comprised most of the species richness detected across lakes. this may explain why, in a previous study where we explored the large-scale spatial patterns in the metabolic profiles of boreal lake and river bacterioplankton communities (ruizgonzález et al., 2015b), we found little or no coupling between changes in structure and function. this of course has major implications on our capacity to predict how variations in the observed taxonomic composition may impact community metabolism or their responses to environmental perturbations, and calls for the need to partition assemblages into components that are subjected to different controls, in order to explore the links between the environment, taxonomic identity, and community functioning in these freshwater microbial assemblages. here we aimed at exploring the fate, the environmental drivers and the metabolic profiles of the dispersed bacterial populations along a hydrologic gradient. to do so, using illumina sequencing of the 16s rrna gene, we assessed the taxonomic composition of bacterioplankton communities inhabiting 10 different rivers from two temperate regions in québec, ranging from small headwater streams (strahler order 0 or 1) to large rivers (strahler order 7). the selected rivers flow through very different landscapes, including forested areas, wetlands and scrublands (berggren and del giorgio, 2015), which had been previously shown to influence the patterns of microbial metabolism in these same rivers (berggren and del giorgio, 2015). we searched for two types of dynamics among headwater bacterial taxa: i) those that decreased in abundance along the hydrologic gradient (i.e., ‘tourist’ taxa, sensu newton et al., 2011), which thus represent taxa that no nco mm er cia l u se on ly 224 dispersed bacteria in a river continuum may not be able to thrive in downstream waters, and ii) those that increased in abundance likely due to in-river selection and growth (i.e., ‘seed’ taxa, sensu ruizgonzález et al., 2015a). in order to understand the ecological underpinnings of the two populations, we explored the environmental variables explaining compositional changes of these communities at both extremes of the hydrologic gradient (i.e., where they were dominant), as well as their metabolic potential using biolog ecoplates. we hypothesized that there should also be a watershed-specific signature in the taxonomic composition of the communities dominated by ‘tourist’ taxa, which reflect the broad catchment differences, whereas assemblages dominated by ‘seed’ taxa should reflect in-river physicochemical conditions. methods study sites, sampling and basic parameters the sampling design has been previously described in berggren and del giorgio (2015). briefly, we sampled 10 streams and rivers for the characterization of bacterioplankton communities with illumina sequencing (illumina inc., san diego, ca, usa) of the 16s rrna gene. the streams were located in two regions of quèbec, canada, and spanned the whole hydrologic gradient ranging from small headwater streams of strahler order 0 or 1 to large rivers (order 6 or 7); they were also selected to maximize the differences in catchment type (peat bogs versus a range of forest soils) and total upstream distance of each site (tab. 1). additional details on the environmental characteristics of the sites can be found in tab. 1 of berggren and del giorgio (2015). each site was sampled twice between 31 may and 24 august 2010, during late spring and summer low-flow conditions, in order to assess the consistency of the patterns observed. two of the rivers could be sampled only once (tab. 1). water was collected at depths of 20-30 cm and all samples were stored in cooling boxes in the dark until analysis. water temperature, dissolved oxygen (do), ph and conductivity were measured in situ with an ysi probe. at each site, a water sub-sample was filtered in situ through 0.45 µm (pes cartridge, sarstedt) and stored in acid-washed glass vials for dissolved organic carbon (doc), nutrients and optical analyses, and another sample was kept in acid-rinsed bottles for dna processing in the laboratory and to measure the substrate utilization profiles using biolog ecoplatestm (see below). doc concentration was measured on an oi1010 toc analyser. the optical properties of doc were measured as indices of its composition and source. coloured dissolved organic matter (cdom) was quantified as the absorbance at 440 nm using an ultrospec 3100 spectrophotometer. doc composition was described on the basis of fluorescence absorption/emission spectra (eems) measured in a shimadzu rf5301 pc spectrofluorophotometer across excitation/emission wavelengths of 275-450 nm and 280600 nm, respectively. six main fluorescence components were recovered from the eems using parallel factor analysis (parafac): components 1 to 3 (c1-c3), previously related to refractory, humic material of presumably terrestrial origin, and c4-c6, associated with more biolabile freshly produced dom (for details see lapierre and del giorgio, 2014; stubbins et al., 2014). c5 and c6 have been linked to autochthonous processes, having c6 a fluorescence signature typical of protein-like material. the percentage contribution of each component was calculated relative to the total fluorescence of the six parafac components. geographical analyses river length, total upstream distance and strahler order, catchment areas and elevation of the sampled sites were obtained using the arcmap 10 and arcgis v10 tab. 1. characteristics of the different sampling sites ordered based on their distance to their respective headwaters. most rivers were sampled in two occasions, but two sites were sampled only once. name region season strahler river order upstream distance (m) catchment type stream # 09 abitibi spring/summer 0 454 broad-leaf forest stream # 08 abitibi spring/summer 1 1193 broad-leaf forest stream # 07 abitibi spring/summer 1 1828 broad-leaf forest petit nord suivit baie james spring/summer 2 2205 bog petit aiguebelle suivit abitibi spring/summer 2 2951 mixed forest ruisseau brunet abitibi spring/summer 4 7149 mixed forest inter-nord suivit baie james spring 4 19575 bog magousi abitibi summer 6 66488 clay soils kinojevis abitibi spring/summer 7 172271 clay soils harricana baie james spring/summer 7 232654 bog no nco mm er cia l u se on ly 225c. ruiz-gonzález et al. software (esri inc., redland, ca, usa) applied on digital elevation models (dem) derived from (1:50 000) maps. for each catchment, we averaged slope based on dem, and determined various landcover properties from geobase (2009). bacterial community taxonomic composition bacterial biomass was collected filtering 300–500 ml of water through 0.22 mm (47 mm, gswp, millipore). bacterial dna was extracted using the mobio (carlsbad, ca, usa) dna extraction kit, following the manufacturer’s instructions. tagged amplicons of the 16srrna gene (v3-v4 region) were obtained with the primers 515f and 806r using a two-step pcr and the access array barcode library (fluidigm), and sequenced on an illumina miseq2000 using a paired-end approach (caporaso et al., 2012). paired-end reads were assembled with flash (magoc and salzberg, 2011) and sequences between 250 and 290 bp were used for downstream analyses in qiime to remove primers, low-quality, archaeal and chloroplast reads (caporaso et al., 2010). quality sequences were binned into operational taxonomic units (otus, ≥97% similarity) using uclust v1.22q (edgar, 2010) and rdp classifier (wang et al., 2007). representative sequences were then aligned against the silvav108 reference alignment (pruesse et al., 2007). to ensure that rare bacteria were not the result of sequencing errors, we discarded all otus showing <4 sequences. to enable comparisons between samples, the otu table was randomly subsampled to ensure an equal number of sequences per sample, based on the sample with the least number of reads (72,798 sequences). identification of ‘seed’ and ‘tourist’ otus in order to explore the dynamics of the taxa dispersed along the hydrologic continuum, we first selected all the otus present in the smallest streams of strahler order 0 (those that did not appear in topographic maps because they are too small or temporary) or order 1. among those, we further categorized otus as i) ‘tourists’ (sensu newton et al., 2011), if their abundances were higher in streams than in the largest rivers of order 6 or 7 (which may represent terrestrial taxa washed from soils that cannot thrive in aquatic conditions), or ii) ‘seeds’, if their abundances were higher in the largest rivers than in headwaters of order 0 and 1 (and thus might represent taxa seeding aquatic communities; ruiz-gonzález et al., 2015a). bacterial substrate utilization patterns the capacity of the riverine bacterial communities to respire 31 carbon substrates was assessed using the biolog ecoplatetm (ca, usa) as detailed in berggren and del giorgio (2015). briefly, one plate per site was incubated in the dark and at 20°c, and the absorbance was measured periodically (every 6 or 12 hours depending on colour development rates) until an asymptote was reached, between 2 and 9 days. at each time point, the overall colour development of the plates was expressed as the average well colour development (awcd), and the mean colour development of each compound was calculated as the blankcorrected mean absorbance of each substrate measured at the time when the awcd was closest to 0.5 (garland et al., 2001), usually between days 1 and 4. all biolog incubations were initiated 2-6 h after field sampling. statistical analyses differences in bacterial taxonomic composition between ecosystem types were tested with anosim (clarke, 1993). the shannon index was calculated as an estimate of bacterial taxonomic diversity. bray-curtis distance was used as an estimator of taxonomic dissimilarity between samples, and was visualised by non-metric multidimensional scaling (nmds) analysis. we performed least squares linear regression to establish relationships between individual variables. all analyses were performed in jmp 9.0.1 (sas institute, nc, usa) or r 3.0.0 software (r core team, 2013; vegan package). results after rarefaction, sequencing of the 16s rrna gene from the 18 samples resulted in 1,310,364 quality sequences, which clustered into 64,489 otus. altogether, 59 bacterial phyla were detected, yet sequences were mainly dominated by the phylum proteobacteria (45% of the total number of sequences), followed by actinobacteria (14%), bacteroidetes (12%), and od1 (7%) and verrucomicrobia (4%). the number of otus per site ranged from 4376 to 17604, and we observed pronounced decreases in otu richness and taxonomic diversity (i.e., shannon index) along a gradient of increasing distance from the headwaters (fig. 1 a,b). interestingly, this pattern was also reflected in a clear structuring of communities depending on the distance travelled (fig. 1f) regardless of the season, region, or catchment type (figs 1 c-e). in order to explore the spatial dynamics of the taxa dispersed along the hydrologic continuum, we identified, among the otus present at the origin of the hydrologic gradient (i.e., in the smallest headwaters), two groups of taxa: ‘tourists’, if their abundances decreased from headwaters to large rivers, or ‘seeds’ if their abundances increased towards downstream waters (see methods). we assume that whereas the former represent taxa washed from soils that cannot thrive in the aquatic environment, the latter comprise taxa (presumably also washed from soils, see ruiz-gonzález et al., 2015a) able to thrive in the aquatic no nco mm er cia l u se on ly 226 dispersed bacteria in a river continuum environment. doing so, 38,725 otus were categorized as ‘tourists’, and only 6026 as ‘seeds’ (fig. 2). together, these two groups represented 92.4% of the total sequences in the dataset. the remaining taxa, 19,738 otus, could not be clearly categorized in any of the two groups due to extreme rareness and lack of any discernible spatial pattern, but these uncategorized taxa represented a very small fraction of the total sequences (7.4%), meaning that these simple criteria recovered most of the diversity across the sampling sites. we observed clear contrasting patterns between the two pools of taxa (fig. 2): ‘tourist’ otus largely dominated headwater communities in all cases but showed pronounced decreases towards the largest rivers (fig. 2a), whereas ‘seed’ otus shifted from making up less than 9% of headwater sequences to numerically dominate the largest rivers (ca. 80% of sequences in rivers of order 6 and 7, fig. 2b). the decrease in ‘tourist’ otu abundance was clearly due to a comparable reduction in the number of otus (fig. 2c), but the dominance of ‘seeds’ in downstream ecosystems seemed caused by large increases in abundance of a small number of otus dispersed from the headwaters (fig. 2d). interestingly, in both cases we observed that such shifts in abundance stabilized after a cumulative distance of around 7 km from the headwaters, which corresponded to rivers of strahler order 4 (fig. 2). beyond that point, both the ‘tourist’ and ‘seed’ otus did not show further changes in relative abundance regardless of the very large differences in the total upstream distance. in terms of the taxonomic composition, the two taxa pools differed largely from each other (fig. 3), and whereas downstream rivers were numerically dominated by groups like betaproteobacteria, actinobacteria and cyanobacteria, in headwater streams we found higher abundances of alpha-, deltaand gammaproteobacteria, od1 or op3, among others (fig. 3). interestingly, this distribution of taxonomic groups was remarkably similar between the two sampling occasions. when the two taxa pools were visualized separately in an nmds, we found that in both cases the communities were also structured following the patterns as a function of distance from the headwaters described above for the whole community (fig. 4), and again this pattern was independent of the season, region or catchment type (details not shown). interestingly, moreover, we observed that, in those sites where either ‘tourist’ or ‘seed’ otus were most abundant, we could detect two different groups of communities based on their taxonomic composition (t1 and t2 in ‘tourist’-dominated communities and s1 and s2 in ‘seed’-dominated communities, see arrows in fig. 4). fig. 1. change in otu number (a) and taxonomic diversity (shannon index, b) as a function of the measured distance between each sampling point and its headwaters. the lines are a best-fit smooth curve through the centre of the data calculated using weighted least squares. each pair of dots represents two sampling points (spring, black; summer, grey) of the same site. c-f) non-metric multidimensional scaling (nmds) plots based on bray-–curtis distances of taxonomic composition of all communities, color-coded by season (c), catchment type (d), region (e) and distance from the headwaters (f). the size of the dot is proportional the number of otus per community. no nco mm er cia l u se on ly 227c. ruiz-gonzález et al. fig. 3. taxonomic composition of ‘seed’ and ‘tourist’ otus along a gradient of increasing distance from the headwaters. data are shown separately for the two sampling times, and are presented as the sum of total sequences associated to each group. the classification was performed at the phylum level in most cases, but the phyla proteobacteria (prot) and bacteroidetes (bact) were split into classes. fig. 2. contribution of headwater otus (i.e., those otus detected in streams of order 0 and 1, see methods) to total bacterial sequences (a,c) or otus (c,d) in aquatic communities as a function of the measured distance between each sampling point and its headwaters. headwater otus were divided depending on their dynamics along the hydrologic continuum into ‘tourist’ otus (otus whose abundances decreased towards rivers, a,c) and ‘seed’ otus (otus whose abundances increased towards rivers, b,d). the lines are a bestfit smooth curve through the center of the data calculated using weighted least squares. each pair of dots represents two sampling points (spring, summer) of the same site. no nco mm er cia l u se on ly 228 dispersed bacteria in a river continuum it is important to note that these two groups of sites did not differ largely in the relative abundance of either ‘seeds’ or ‘tourist’ otus regardless of the differences in distance travelled (figs. 2 and 4). therefore, the observed changes in composition between t1 and t2, and between s1 and s2, must be due to changes on the group membership as a response to environmental factors and not only to direct hydrologic transport. in order to assess so, using envfit analysis we explored the environmental variables that best explained the taxonomic differences between those two groups of sites (t1 vs t2 and s1 vs s2) for each of the taxa pools (i.e., the smallest headwater streams in the case of communities dominated by ‘tourist’ otus, and the largest rivers in the case of communities dominated by ‘seed’ otus, fig. 5). we observed that the compositional differences between the two groups of communities dominated by ‘tourist’ otus (t1 and t2) were associated with differences in the ratio of cdom to doc, which is a proxy of terrestrial influence, as well as in the proportions of the dom fluorescent components c1 (associated to processed humic-like material) and c3 (freshly produced humic-like material, fig. 5a). conversely, variation between the two groups of ‘seed’-dominated communities in the largest rivers (s1 and s2) seemed to be due to differences in autochthonous processes, since chlorophyll a concentration and % of c5 (associated to algal material) component appeared as the variables most clearly explaining the observed pattern, together with another humic-like component c2 (fig. 5b). in general, the taxonomic differences between ‘seed’ dominated communities were larger than those between communities dominated by ‘tourist’ otus (see axes in fig. 5). finally, we investigated differences between the metabolic profiles (assessed by the biolog ecoplates) of all communities (fig. 6). interestingly, we observed clear metabolic differences between communities dominated by ‘seed’ and ‘tourist’ otus, with the exception of one site that corresponded to the only river of order 6, which grouped together with the headwater streams (fig. 6a). these differences seemed mostly caused by variations in the use of four carbohydrates (d-cellobiose, d-lactose, glucose, erythritol), one carboxylic acid (2-hydroxy benzoic acid) and one amino acid (l-serine, fig. 6a). in general, communities dominated by ‘seed’ otus (i.e., large rivers) displayed higher capacity to respire l-serine and erythritol and lower capacity to use d-cellobiose, dlactose and 2-hydroxy benzoic acid than communities dominated by ‘tourist’ otus. however, ‘tourist’-dominated communities showed larger metabolic differences between groups t1 and t2 than ‘seed’ communities, which showed less clear differentiation between s1 and s2 (fig. 6a). interestingly, we observed a gradient of increasing metabolic dissimilarity along a gradient of river order (fig. 6b), suggesting that communities from the small headwater streams were more functionally similar to each other than communities located towards the lower end of the hydrologic continuum. fig. 4. non-metric multidimensional scaling (nmds) plots based on bray-curtis distances of taxonomic composition of both the ‘seeds’ and ‘tourist’ otus (see legend of fig. 1). the color gradient indicates the distance between each sampling point and its headwaters, and the size of the dot is proportional to the relative contribution of either ‘seed’ or ‘tourist’ otus to each community’ sequences. for each taxa pool, the two arrows indicate the two groups of communities where each type of otu dominated (t1 and t2 in ‘tourist’dominated communities and s1 and s2 in ‘seed’-dominated communities), which were compared in order to understand the environmental drivers explaining the observed changes in composition but not in abundance (see figs. 5 and 6). no nco mm er cia l u se on ly 229c. ruiz-gonzález et al. discussion our results show a sequential structuring of riverine bacterioplankton communities, which changed from highly diverse communities inhabiting the smallest headwater streams, towards assemblages with much lower taxonomic richness and diversity inhabiting the largest rivers (fig. 1 ,b). this pattern was also reflected by an organization of communities along a gradient of cumulative distance from the headwaters (i.e., travelled distance, fig. 1f) that coincided with concomitant increases in river order (details not shown) and likely also water residence time, although the later was not directly measured. the fact that the observed structure was independent from the geographic location of the rivers, the sampling period, and the catchment type (fig. 1), supports previous results showing a major role of hydrology and landscape connectivity in shaping the structure of freshwater bacterioplankton communities (crump et al., 2007; nelson et al., 2009; niño-garcía et al., 2016a). the high numbers of otus that we observed in the smallest headwater streams agrees with the notion that these systems act as integrators and dispersers of microbial diversity originating from the surrounding terrestrial landscape (savio et al., 2015; hauptmann et al., 2016; ruiz-gonzález et al., 2015a; niño-garcía et al 2016a). moreover, whereas the observed decrease in otu richness suggests that many of these dispersed bacteria cannot thrive in the water, the reduction in taxonomic diversity further implies that there is a progressive dominance of taxa adapted to aquatic conditions, as previously reported (savio et al., 2015; niño-garcía et al 2016a). in order to check whether we could differentiate such contrasting dynamics within the dispersed taxa, we identified those otus present at the origin of the hydrologic continuum (i.e., streams of order 0 or 1) that i) decreased in abundance (i.e., ‘tourist’ otus) or ii) increased in abundance (‘seed’ otus) towards the largest rivers. doing so, 58% of all otus were categorized as ‘tourist’, and only 9% as ‘seeds’ (fig. 2), but together represented 92.4% of the total sequences in the dataset, suggesting that these simple criteria captured the most important community dynamics across the studied riverine ecosystems. we observed pronounced decreases in otu number and relative contribution of ‘tourist’ taxa from the headwaters to the largest rivers. these two parameters (‘tourist’ otu richness and abundance, fig. 2 a,c) showed a very similar pattern and indeed largely covaried (r2=0.93, n=18, p<0.0001), suggesting that the observed abundance decrease of ‘tourist’ otus was likely the result of dilution or removal of taxa that could not grow in the aquatic environment. the presence of typical soil or groundwater groups such as acidobacteria, deltaproteobacteria, thermi, od1, op3, verrucomicrobia within the ‘tourist’ group supports the fact that their presence is largely due to advection from the catchment (janssen et al., 2002; lauber et al., 2009; barberán et al., 2014; hiller et al., 2015). conversely, ‘seed’ otus displayed pronounced increases in abundance along the hydrologic gradient that were not accompanied by comparable increases in the number of otus (fig. 2 fig. 5. pca analysis of the communities where ‘seed’ (a) and ‘tourist’ (b) otus dominated, color-coded by the two groups of sites (t1 and t2 in ‘tourist’-dominated communities and s1 and s2 in ‘seed’-dominated communities) observed in fig. 4. the vectors show the environmental variables that best fitted the ordination space (using the r envfit funtion). the size of the vector is proportional to the strength of the correlation of each variable. cdom/doc, ratio of coloured dissolved organic matter to dissolved organic carbon concentration (a proxy of terrestrial inputs); chla, chlorophyll a concentration; %c1, %c2, %c3,%c5, percentage contribution of the fluorescent parallel factor analysis components c1,2,3 and c5 to the total fluorescence. no nco mm er cia l u se on ly 230 dispersed bacteria in a river continuum b,d). this indicates that ‘seed’ otus comprise a smaller pool of taxa, also present at the origin of the hydrologic continuum, which grow during transit through riverine networks and dominate downstream assemblages. based on previous studies, it is likely that most of these ‘seeds’ are actually washed from the surrounding soils, where they can persist at low abundances (crump et al., 2012; ruizgonzález et al., 2015a), maybe dormant (lennon and jones 2011). moreover, this distinction between ‘seeds’ and ‘tourist’ otus is in accordance with the partition of boreal lake bacterioplankton communities into ‘core’ (adapted to local conditions) and ‘accidental’ (linked to hydrologic transport) components based on their large-scale spatial dynamics (niño-garcía et al., 2016b), and supports the notion that the main community assembly mechanisms change gradually along hydrologic gradients, shifting from communities strongly subjected to mass effects in the fast flowing headwater streams that are highly influenced by terrestrial inputs, to assemblages where local sorting of species will prevail (lindström et al., 2006; crump et al., 2007; nelson et al., 2009; ruiz-gonzález et al.,, 2015; niño-garcía et al., 2016a). interestingly, we observed that such increases and decreases in the abundances of ‘seed’ and ‘tourist’ otus, respectively, tended to stabilize beyond a travelled distance of ca. 7 km (fig. 2 a,b), corresponding to the rivers of order 4. this coincides with our own previous observation, in a different region of boreal québec, that the proportion of sequences associated to the soil-derived taxa able to grow in the water increased progressively from the smallest streams until rivers of order 4, beyond which their contribution to total sequences tended to plateau at ca. 70% (ruiz-gonzález et al., 2015a). since this stabilization corresponded approximately to a water residence time of 10 days, it was suggested that beyond 10 days hydrology no longer constrains the growth of taxa, and thus any further variations in community composition will be due to local selection driven by changes in aquatic physicochemical or biological conditions (ruiz-gonzález et al., 2015a; niño-garcía et al., 2016a). accordingly, here we observed that although beyond 7 km the relative contribution of ‘seed’ otus remained remarkably constant (around 80% of the community sequences), there were pronounced changes in taxonomic composition between those communities (likely due to changes in river physicochemical conditions), which showed two clear groups of sites corresponding to rivers of order 4 (s1) and rivers of order 6-7 (s2, fig. 4b). similarly, fig. 6. a) pca based on the metabolic profiles (biolog ecoplates) of all communities. different shapes indicate whether ‘tourist’ or ‘seed’ otus dominated the communities (see arrows in fig. 4). the vectors represent the carbon substrates that significantly explained the observed ordination; ca, carboxylic acid; ch, crbohydrate; aa, amino acid. b) increases in metabolic dissimilarity (assessed as euclidean distance between samples characterized through biolog profiles) as a function of strahler river order. no nco mm er cia l u se on ly 231c. ruiz-gonzález et al. at the other end of the hydrologic gradient, communities dominated by ‘tourist’ otus (i.e., those in the smallest headwater streams, order 0-2) also clustered into two different groups of sites (t1 and t2, fig. 4a). we expected that such compositional differences in community composition would be driven by different environmental factors at each extreme of the hydrologic gradient, since ‘tourist’ taxa should mainly reflect the terrestrial environments where they originate, whereas ‘seed’ taxa should be subjected to species selection driven by local aquatic environmental conditions. to assess so, we split the dataset into communities dominated by ‘tourist’ or ‘seed’ otus, respectively, and explored their variability in relation to all the measured variables (fig. 5). variables associated to dom quality emerged as the most strongly related to the ordination in both cases, but they differed between ‘seed’and ‘tourist’dominated communities (fig. 5 a,b). for example, taxonomic differences between the two groups of communities dominated by ‘tourist’ taxa in streams (t1 and t2) seemed to be related to differences in the ratio of cdom to doc (a proxy of terrestrial influence; lapierre and del giorgio, 2014) or in the proportion of the dom fluorescent components c1 (associated to processed humic-like material) and c3 (freshly produced humic-like material, fig. 5a). although we could not see a clear clustering of communities based on the different catchment types shown in tab. 1, not even for the headwater assemblages (fig. 1), this observation supports the hypothesis that headwater streams reflect to some extent a signature of the surrounding catchment, which in these same rivers had been shown to provide doc of different quality (berggren and del giorgio, 2015). in this regard, niño-garcía et al., (2016) observed a strong structuring effect of ph on boreal bacterioplankton communities, but attributed this to a regional landscape signature on the immigrant pool rather than to species sorting due to ph itself. similarly, lear et al., (2013) found that the biogeography of stream biofilm bacterial communities was related to variability in catchment land use, supporting a strong influence of the surrounding terrestrial environment in systems with high connectivity with land. on the other hand, taxonomic variation across ‘seed’ communities in the largest rivers appeared to be driven by differences in local aquatic processes, since chlorophyll a concentration and %c5 (associated to freshly-produced algal material) component appeared as significant environmental variables (fig. 5b) and varied largely between the two groups of sites s1 and s2. chlorophyll concentrations were higher in larger rivers independently of the surrounding catchment type (details not shown), which agrees with the expected pattern of downstream increase in planktonic chlorophyll, associated with longer residence times and increases in nutrients (sabater et al., 2008), and this likely triggered the growth of bacterial taxa adapted to more productive conditions in s2 than in s1 sites. in any case, the observation that all these ‘seed’ taxa, regardless of their taxonomic identity, were detected in the headwater streams supports that landscapes in these regions contain a reservoir of diversity able to colonize aquatic ecosystems if the water residence time is long enough (ruiz-gonzález et al., 2015a). along this same line, it is remarkable to note that the studied streams and rivers were not physically connected; all of them were located in different watersheds in two different geographic regions, separated by more than 350 km. therefore, the fact that those otus present across the sampled headwaters included taxa that would become dominant in all of the larger rivers suggests that there must be a large regional pool of taxa from which aquatic communities recruit, which is not subjected to dispersal limitation. interestingly, this idea of the landscape reservoir of aquatic diversity has been proposed at scales much larger than the present study (niño-garcía et al., 2016a; ruiz-gonzález et al., 2015a, 2017) and is in accordance with the seed-bank concept, a standing reservoir of dormant microbes (i.e., microbes in a reversible state of very low metabolic activity) that persist at low abundances but that can grow and become abundant upon changes in environmental conditions (lennon and jones, 2011). this further highlights that microbial seed banks may transcend ecosystem types, such that viable microbes can persist under unfavourable conditions in some ecosystems until they are transported to a suitable environment (sjöstedt et al., 2012; lee et al., 2013; comte et al., 2014; ruiz-gonzález et al., 2015a, 2017). soils in particular are known to contain a large fraction of dormant cells (lennon and jones, 2011), a fraction of which seems able to resuscitate and grow upon advection to the water. in any case, this fact challenges our current ecosystem-specific approaches to study bacterial communities, and highlights that it will not be possible to understand how communities are assembled or structured without considering at least the most important potential sources of microbial dispersal. finally, the fact that such a large fraction of the diversity may represent taxa that are not adapted to local conditions has important implications for our understanding of community functioning. this agrees with recent evidence suggesting that a potentially large fraction of the microbial diversity detected in biogeographic studies can be either dead, in a dormant state, or present as extracellular dna (lennon and jones, 2011; blagodatskaya and kuzyakov, 2013; carini et al., 2016). in this regard, exploring the dynamics of different taxa across ca. 200 lakes niño-garcía et al. (2016a) revealed that the presence of most rare bacterial taxa was due to hydrologic transport from the associated networks, and thus these taxa likely represent a species pool that is not contributing to the functioning of communities but that is detected in all sequencing studies. for these reasons, we expected to find no nco mm er cia l u se on ly 232 dispersed bacteria in a river continuum a clearer link between taxonomy and the metabolic profiles of the studied communities when considering separately the pool of ‘seed’ and ‘tourist’ taxa than when considering the whole communities, and in addition, we expected to find a stronger correlation between taxonomy and function for ‘seeds’ than for ‘tourist’ otus. however, this was not the case: changes in taxonomic composition were weakly but significantly correlated to changes in metabolic profiles in all three cases, i.e., when the whole community was considered (mantel r=0.189, p<0.05), or when considering ‘seeds’ (mantel r=0.201, p<0.05) or ‘tourist’ otus (mantel r=0.215, p<0.05) separately, and the strength of the correlation was very similar in all cases. this may be due to the fact that the biolog ecoplates used for assessing the metabolic profiles do not measure in situ activity levels but rather a functional potential, because it is a method that relies on the growth of microorganisms on the selected substrates during an incubation of several days (garland et al., 2001). thus, the microorganisms that will grow on the plates may be rare in the original communities (smalla et al., 1998), particularly in headwaters where taxa do not have the time to attain high abundances due to the fast hydraulics. however, we did observe clear metabolic differences between communities dominated by ‘seeds’ and ‘tourist’ otus (fig. 6a), suggesting that the functional structure or potential of these communities actually differs. interestingly, moreover, whereas communities from the headwater streams were functionally more similar to each other, assemblages in large rivers seemed much more heterogeneous metabolically (fig. 6b). this agrees with a previous study where we analyzed the metabolic profiles of ca. 300 freshwater bacterioplankton assemblages and found that communities from systems more tightly connected to the land were more similar metabolically than communities from larger lakes or rivers, and this increase in metabolic dissimilarity followed a gradient of increasing concentrations of autochthonously produced organic carbon (ruizgonzález et al., 2015b). this pattern may suggest that as systems depart from land, the structural diversification of the available dom pool due to local processes (kothawala et al., 2014) causes a functional diversification, whereas terrestrial dom is more homogeneous or simply the pool of headwater taxa that can grow on the substrates is more restricted that in true aquatic communities. it is important to note that this diversification of function along the aquatic flow path does not come in parallel to a taxonomic diversification; in fact, there was an overall decline in otu richness along this same continuum. in a previous study on these same rivers (berggren and del giorgio, 2015) the type of catchment (mainly forest headwater versus peat influenced) was shown to be the strongest determinant of multiple facets of bacterial metabolism (e.g., biomass production, respiration, growth efficiency) presumably driven by watershed-scale differences in doc quality and quantity. indeed, we found that catchment type was the variable explaining most of the metabolic differences between communities (tab. 2). however, we also found weaker but significant differences based on the four types of communities (tab. 2), suggesting that the taxonomic composition also plays a role in determining the metabolic potential of communities in these heterogeneous rivers. conclusions overall, our results show that bacterioplankton communities from lotic systems may comprise varying proportions of taxa adapted and not adapted to live in aquatic conditions, and this balance is largely driven by the time that bacteria have spent in a given parcel of water. since the assembly mechanisms of both components differ, the environmental variables explaining shifts in taxonomic composition also change between the two pools of taxa, and whereas variables related with terrestrial inputs from the catchment explained differences in communities dominated by ‘tourist’ otus, varying concentrations of chlorophyll and the associated doc appeared to be related to compositional changes in ‘seed’-dominated communities. finally, we found that the metabolic potential associated to these contrasting taxa pools also differed. all this highlights the need to consider the landscape connectivity between assemblages and the resulting dynamics of specific populations in order to fully understand the assembly mechanisms and environmental factors shaping these flowing communities. acknowledgements we thank the whole carbbas team for their contribution to the field and laboratory components of this retab. 2. anosim r values indicating significant differences in biolog metabolic profiles between different categories. all r values for euclidean distance matrices were calculated using 9999 permutations. significance codes for r values community group refers to the 4 groups of communities where either ‘seed’ or ‘tourist’ otus dominated (see arrows in fig. 4). anosim r values by season ns by region ns by catchment type 038** by river order 0.19* by community group 0.16* ns, not significant; **p<0.0001; *p<0.05. no nco mm er cia l u se on ly 233c. ruiz-gonzález et al. search. in particular, we thank j.f. lapierre for providing the parafac data. this study is part of the program of the carbon biogeochemistry in boreal aquatic systems (carbbas) industrial research chair, co-funded by the natural science and engineering research council of canada (nserc) and hydro-quebec. we acknowledge funding from frqnt, nserc, and the canada research chairs program. crg benefited of a juan de la cierva post-doctoral fellowship from the spanish ministry of industry, economy and competitivity. references barberán a, ramirez ks, left jw, bradford ma, wall dh, fierer n, 2014. why are some microbes more ubiquitous than others? predicting the habitat breadth of soil bacteria. ecol. lett. 17:794-802. berggren m, del giorgio pa, 2015. distinct patterns of microbial metabolism associated to riverine dissolved organic carbon of different source and quality. j. geophys. res. biogeo.120:989-999. besemer k, singer g, quince c, bertuzzo e, sloan w, battin tj 2013. headwaters are critical reservoirs of microbial diversity for fluvial networks. p. roy. soc. b. 280:20131760. besemer k, peter h, logue jb, langenheder s, lindstrom es, tranvik lj, battin tj, 2012. unraveling assembly of stream biofilm communities. isme j. 6:1459-1468. blagodatskaya e, kuzyakov y, 2013. active microorganisms in soil: critical review of estimation criteria and approaches. soil biol. biochem. 67:192-211. caporaso jg, lauber cl, walters wa, berg-lyons d, huntley j, fierer n, et al. 2012. ultra-high-throughput microbial community analysis on the illumina hiseq and miseq platforms. isme j. 6:1621-1624. caporaso jg, kuczynski j, stombaugh j, bittinger k, bushman fd, costello ek, et al. 2010. qiime allows analysis of high-throughput community sequencing data. nat. methods 7:335-336. carini p, marsden pj, leff jw, morgan ee, strickland ms, fierer n, 2016. relic dna is abundant in soil and obscures estimates of soil microbial diversity. nat. microbiol. 2:16242. clarke kr, 1993. non-parametric multivariate analysis of changes in community structure. australian j. ecol. 18:117-143. comte j, lindström es, eiler a, langenheder s, 2014. can marine bacteria be recruited from freshwater sources and the air? isme j. 8:2423-2430. crump bc, amaral-zettler la, kling gw, 2012. microbial diversity in arctic freshwaters is structured by inoculation of microbes from soils. isme j. 6:1629-1639. crump bc, adams he, hobbie je, kling gw, 2007. biogeography of bacterioplankton in lakes and streams of an arctic tundra catchment. ecology 88:1365-1378. edgar rc, 2010. search and clustering orders of magnitude faster than blast. bioinformatics 26:2460-2461. garland jl, mills al, young js, 2001. relative effectiveness of kinetic analysis vs single point readings for classifying environmental samples based on community-level physiological profiles clpp. soil biol. biochem. 33:1059-1066. hiller ka, foreman kh, weisman d, bowen jl, 2015. permeable reactive barriers designed to mitigate eutrophication alter bacterial community composition and aquifer redox conditions. appl. environ. microbiol. 81:7114-7124. janssen ph, yates ps, grinton be, taylor pm, sait m, 2002. improved culturability of soil bacteria and isolation in pure culture of novel members of the divisions acidobacteria, actinobacteria, proteobacteria, and verrucomicrobia. appl. environ. microbiol. 68:2391-2396. kothawala dn, stedmon ca, müller ra, weyhenmeyer ga, köhler sj, tranvik l, 2014. controls of dissolved organic matter quality:evidence from a large-scale boreal lake survey. glob. change biol. 20:1101-1114. lapierre jf, del giorgio pa, 2014. partial coupling and differential regulation of biologically and photochemically labile dissolved organic carbon across boreal aquatic networks. biogeosciences 11:5969-5985. lauber cl, hamady m, knight r, fierer n, 2009. pyrosequencing-based assessment of soil ph as a predictor of soil bacterial community structure at the continental scale. appl. environ. microbiol. 75:5111-5120. lear g, washington v, neale m, case b, buckley h, lewis g, 2013. the biogeography of stream bacteria. global ecol. biogeogr. 22:544-554. lee je, buckley hl, etienne rs, lear g, 2013. both species sorting and neutral processes drive assembly of bacterial communities in aquatic microcosms. fems microbiol. ecol. 86:288-302. lennon jt, jones se, 2011. microbial seed banks:the ecological and evolutionary implications of dormancy. nat. rev. microbiol. 119:119-130. lindström es, bergström ak, 2004. influence of inlet bacteria on bacterioplankton assemblage composition in lakes of different hydraulic retention time. limnol. oceanogr. 49:125-136. lindström es, forslund m, algesten g, bergström a, 2006. external control of bacterial community structure in lakes. limnol. oceanogr. 51:339-342. logares r, lindström es, langenheder s, logue jb, paterson h, laybourn-parry j, et al. 2013. biogeography of bacterial communities exposed to progressive long-term environmental change. isme j. 7:937-948. magoc t, salzberg s, 2011. flash: fast length adjustment of short reads to improve genome assemblies. bioinformatics 27:2957-2963. nelson ce, sadro s, melack jm, 2009. contrasting the influences of stream inputs and landscape position on bacterioplankton community structure and dissolved organic matter composition in high-elevation lake chains. limnol. oceanogr. 54:1292-1305. newton rj, jones se, eiler a, mcmahon kd, bertilsson s, 2011. a guide to the natural history of freshwater lake bacteria. microbiol. mol. biol. res. 75:14-49. niño-garcía jp, ruiz-gonzález c, del giorgio pa, 2016a. interactions between hydrology and water chemistry shape bacterioplankton biogeography across boreal freshwater networks. isme j. 10:1755-1766. niño-garcía jp, ruiz-gonzález c, del giorgio pa, 2016b. landscape-scale spatial abundance distributions discriminate core from random components of boreal lake bacterioplankton. ecol. lett. 19:1506-1515. no nco mm er cia l u se on ly 234 dispersed bacteria in a river continuum pruesse e, quast c, knittel k, fuchs bm, ludwig w, peplies j, et al. 2007. silva:a comprehensive online resource for quality checked and aligned ribosomal rna sequence data compatible with arb. nucleic acids res. 35:71887196. ruiz-gonzález c, niño-garcía jp, del giorgio pa, 2015a. terrestrial origin of bacterial communities in complex boreal freshwater networks. ecol. lett. 18:1198-1206. ruiz-gonzález c, niño-garcía jp, lapierre jf, del giorgio pa, 2015b. the quality of organic matter shapes the functional biogeography of bacterioplankton across boreal freshwater ecosystems. global ecol. biogeogr. 11:11-12. ruiz-gonzález c, niño-garcía jp, kembel sw, del giorgio pa, 2017. identifying the core seed bank of a complex boreal bacterial metacommunity. isme j. 11:2012-2021. savio d, sinclair l, ijaz uz, et al., 2015. bacterial diversity along a 2600 km river continuum. environ. microbiol. 17:4994-5007. sjöstedt j, koch-schmidt p, pontarp m, canbäck b, tunlid a, lundberg p, et al. 2012. recruitment of members from the rare biosphere of marine bacterioplankton communities after an environmental disturbance. appl. environ. microbiol. 78:1361-1369. smalla k, watchtendorf u, heuer h, liu w, forney l, 1998. analysis of biolog gn substrate utilization patterns by microbial communities. appl. environ. microbiol. 64:1220-1225. souffreau c, van der gucht k, van gremberghe i, kosten s, lacerot g, lobão lm, et al., 2015. environmental rather than spatial factors structure bacterioplankton communities in shallow lakes along a >6000 km latitudinal gradient in south america. environ. microbiol. 17:2336-2351. stubbins a, lapierre jf, berggren m, prairie yt, dittmar t, del giorgio pa, 2014. what’s in an eem? molecular signatures associated with dissolved organic fluorescence in boreal canada. envir. sci. tech. 48:10598-10606. wang q, garrity gm, tiedje jm, cole jr, 2007. naïve bayesian classifier for rapid assignment of rrna sequences into the new bacterial taxonomy. appl. environ. microbiol. 73:5261-5267. wilhelm l, besemer k, fragner l, peter h, weckwerth w, battin tj, 2015. altitudinal patterns of diversity and functional traits of metabolically active microorganisms in stream biofilms. isme j. 9:2454-2464. no nco mm er cia l u se on ly layout 1 introduction palaeolimnological tools can be applied to reveal information retained in aquatic sediments (smol 1992, european union 2000). cladocerans (branchiopoda: phyllopoda), which have been demonstrated to be excellent indicators of environmental change (eggermont and martens, 2011), have been used in palaeolimnological studies to reconstruct past environmental conditions of lake ecosystems since the 1950s (frey, 1986). despite their obvious value, methodological limitations and sources of error still exist. cladoceran remains, for example, preserve selectively in lake sediments, as some species and some components preserve better than others. in the case of daphniids o.f. müller, only postabdomal claws and resting spores (ephippia) are typically found in sediments (schmidt et al., 1998; korhola, 1999; sarmajakorjonen, 2002), while the small size of their claws inflicts an elevated risk of losing claws during sieving. this may lead to an underestimation of daphnia o. f. müller abundance when it is based on the number of claws (nykänen et al., 2009). the poor preservation of daphnia is particularly problematic, because daphnia is by far the most studied cladoceran taxon (ebert, 2005). for example, peters and de bernardi (1987) and lampert (2011) published comprehensive reviews of daphnia ecology, and the number of scientific articles on daphnia indexed in the thomson reuters web of science exceeds 10 000 (seda and petrusek, 2011). nevertheless, daphnia remains do preserve well in some lakes, such as in lake kivijärvi (finland), although the reasons behind this variation in preservation are not yet fully understood. the preservation issues in sedimentary cladocerans have been studied before (deevey, 1964; kerfoot, 1981), but we are unaware of previous research attempting to assess the variations in daphnia degradation in a sediment core. in this study we combined data on daphnia preservation to sedimentary geochemistry, diatom-inferred lake water ph, predation indices and known land use and fishing history of lake kivijärvi in order to shed new light on the preservation issues of daphnia remains. methods study area lake kivijärvi is a small (1.8 km2), relatively deep (max depth 12 m), brown-coloured lake, which is surrounded by mires and coniferous forests. it is located at 165 m asl in central northern finland (n 63° 55.998’ e 27° 54.281’). during the summer stratification in august 2006, which corresponds to the age of our topmost sediadvances in oceanography and limnology, 2016; 7(2): 142-150 article doi: 10.4081/aiol.2016.6293 this work is licensed under a creative commons attribution-noncommercial 4.0 international license (cc by-nc 4.0). varying degradation of subfossil daphnia longispina during the past 250 years and the discovery of fossil helmet-type head shields: preliminary results jaakko johannes leppänen,* jan weckström department of environmental sciences, university of helsinki, p.o. box 65, fin-00014, finland *corresponding author jaakko.leppanen@helsinki.fi abstract zooplankton are regarded as a good indicator of environmental change, but comprehensive monitoring programs including zooplankton are uncommon and only rarely extend over longer periods of time. a part of the zooplankton community can be reconstructed using palaeolimnological methods, yet challenges remain. for example, cladoceran subfossil remains preserve selectively in sediments. in particular, the remains of daphnia spp. are known to usually exhibit poor level of preservation; the reasons for this are still unclear. in the rural lake kivijärvi, located in central finland, daphnia subfossil remains preserve extraordinary well and multiple fossil components are found. however, the preservation level is not uniform and exhibits directional change throughout the sediment record. to investigate the changes in daphnia preservation in lake sediments, we graded caudal spines from 20 fossil sediment samples into three taphonomic groups. a dataset of sediment geochemistry, diatom-inferred lake water ph, predation indices, and the catchment land use history was used to assess the environmental history of our study lake. in lake kivijärvi, the most significant change in daphnia preservation seems to correspond best with the historical fishing activities. additional explanatory variables include forestry in the catchment area, and ph, which, however, had contradicting effects on the preservation of daphnia remains in this study. finally, a fossil daphnia longispina helmet type head shield derived from the lake sediment is presented for the first time. key words: daphnia; preservation; subfossil; palaeolimnology; cladocera. received: september 2016. accepted: november 2016. no n c om me rci al us e o nly varying preservation of subfossil daphnia longispina 143 ment sample, the lake surface water (1 m) was slightly acidic (ph 6.2), with an alkalinity of 59 µmol l–1 and conductivity of 19 µs cm–1. the total phosphorus and total nitrogen concentrations measured during this period were 33 µg l–1 and 540 µg l–1, respectively, and the chlorophyll-a concentration 22 µg l–1. the lake is fed by river lumijoki and the brook myllypuro. human activities, such as land use and fishing have affected the lake since the beginning of the 20th century. the catchment area (53 km2) has been heavily ditched during forestry practices, which were most intense during the 1960s and 1970s. in addition, one peat extraction area (33 ha) located in the western part of the kivijärvi catchment operated between 1986 and 2012 (n. huotari, personal communication). today the proportion of the ditched area in the lake catchment is over 40%. more recent disturbances (since 2008) include sulphate and metal pollution from the talvivaara ni-cu-zn-co mine, which is connected to lake kivijärvi via lumijoki river and is located approx. 6 km ne from lake kivijärvi (kauppi et al., 2013). the average (19712000) yearly air temperature and precipitation are +2°c and ~600 mm, respectively (kersalo and pirinen, 2009). annual precipitation has increased by an estimated 0.92 mm±0.50 mm year–1 (irannezhad et al., 2014) during the last century and mean air temperature has increased by 2°c during the past 166 years (mikkonen et al., 2014). bathymetric data was obtained from the finnish environment institute electronic gis library (http://wwwd3.ymparisto.fi/d3/wmsrajapinta.htm), the catchment area was determined using the online tool value provided by the finnish environmental institute (http://paikkatieto.ymparisto.fi/value/) and was analysed using the geographical information system program esri arcmap 10.2.1. water chemistry data was obtained from the oiva database (finnish environment institute, https://wwwp2.ymparisto. fi/scripts/oiva.asp). data regarding forestry activity between 1966 and 1994 was obtained from historical map archives of the national land survey of finland (http://vanhatpainetutkartat.maanmittauslaitos.fi/). coring two short sediment cores were retrieved from the northern section of lake kivijärvi (n 63° 55.998’e 27° 54.281’ wgs84, at the depth of 8 m) using a hth-kajak sediment corer with a diameter of 9 cm (renberg and hansson, 2008). core a (26 cm length) was used for sediment dating, and for analyses of sub-fossil cladocerans and diatoms, core b (24 cm) was used for elemental analysis. the cores were subsampled at 0.25 cm intervals for the first 5 cm and at 1 cm intervals between 5 and 26 cm. the initial reason for different subsampling for the first 5 cm was to provide possibility for high resolution assessment of known mine pollution, which is affecting lake water quality since 2008. mining effluent has resulted in elevated lake metal (fe, mn, ni, zn) and salt (naso) concentrations, and has turned the lake meromictic (kauppi et al., 2013). subsampling was conducted in the field and the samples were stored in plastic zip lock bags and stored in a dark cold room at 4°c. as the two cores were retrieved from the flat and large center of the lake and less than 1.5 m apart from each other, the temporal resolution of these two cores can be assumed as comparable. nevertheless, the cores were correlated based on their visual features. both cores exhibited analogous and clear changes in sediment composition in the top first cm, where the sulfur rich layer originating from the mine pollution disappeared strictly at 0.75-1 cm level. moreover, both cores showed a clear transition layer between 4 and 6 cm, where the sediment color changed to darker brown. below this, both cores showed unchanged sediment color and texture. the top 0-1 cm core samples were omitted from the study, as they were deposited after the onset of mine pollution and therefore strongly affected by metals and sulphur, which caused the almost complete disappearance of daphnia remains. dating and sediment chemistry freeze dried sediment samples from core a were radiometrically dated at the liverpool university environmental laboratory, using ortec hpge gwl well-type coaxial low background intrinsic germanium detectors (appleby et al., 1986). the samples were analysed for 210pb, 226ra, 137cs and 241am by direct gamma assay. elemental concentrations of erosion indicators magnesium (mg), potassium (k) and sodium (na) were analysed in the acid soluble (nho3) sediment fractions in the metropolilab environmental laboratory at helsinki, which is an accredited testing laboratory (finas t058). daphnia analysis each sediment subsample (~1 cm3 of wet material) was treated using 10 % koh solution, heated (80°c) on a hotplate for 30 min, and sieved through 48 µm mesh size with tap water. samples were mounted with safranine stained glycerol. for more information regarding the protocol, see szeroczyñska and sarmaja-korjonen (2007). daphnia remains were identified and counted using both 200 and 400 x magnifications. during the preliminary microscopic screening, a broad variation in the preservation of postabdominal claws was observed, e.g. notable variations in type and magnitude of damage in claws, denticles and setae. in addition, also the size of the claws showed large variability, which, in turn, has a direct impact on their resistance to damage. classification of claws was thus found to be extremely difficult, whereas the differences in degree of preservation were clearer for caudal spines. caudal spines were therefore identified as the most no n c om me rci al us e o nly 144 j.j. leppänen and j. weckström suitable component to study the variance in preservation of subfossil daphnia. to evaluate these preservation differences, 100 spines from each one of the 20 sediment subsamples were identified and assessed for a taphonomic grade. taphonomic grading is used in palaeoecology to assess the deposition speed and reworking of fossil assemblages (brandt, 1989). the studies regarding microbial decomposition of daphnia remains are rare, but it has been reported that daphnia remains are rapidly and heavily colonized by chitin degrading microbes (tang et al., 2009) and fungi (czezuga et al., 2002). the colonization by degrading organisms would imply a rather even and comprehensive degradation within a certain sediment layer than mechanical impacts. in the present work each spine was graded in relation to its stiffness, folding and breakage (tab. 1; fig. 1). we assumed that folded or broken spines have been damaged due to mechanical impact, and spines that have completely lost their form (slack spines) have been affected by microbial or chemical degradation. predation assessment to assess changes in predation on cladocerans in lake kivijärvi, we conducted morphological measurements of eubosmina longispina leydig exoskeletal remains (36-46 measurements of carapace and mucro length per sample), and counted the chaoborus spp. lichtenstein, mandibles detected in the sediment subsamples to calculate the ratio (number of chaoborus spp. mandibles: 100 daphnia spines). the value as an ecological indicator of morphological changes within the genus bosmina has been recently reviewed by korosi et al. (2013), but the applicability of subgenera eubosmina morphology is not so straight forward due to contrasting results in literature (sprules et al., 1984; johnsen and raddum, 1987). in contrast, the abundance of chaoborus larvae has been demonstrated to reflect the abundance of fish and to affect the cladoceran community structure (sweetman and smol, 2006). numerical analysis the non-parametric mann-kendall trend test (gilbert, 1987) was used to detect significant monotonic trends in the taphonomic grade data of sedimentary daphnia remains. constrained optimal sum of squares partitioning with untransformed species percentage data (birks and gordon, 1985) was used to detect significant zones in taphonomic stratigraphy. the number of statistically significant zones was calculated using the broken-stick model described in bennett (1996). optimal partitioning was conducted using the program zone 1.2 (lotter and juggins, 1991). the mann-kendall test was conducted with past statistics 3.10 software (hammer et al., 2001). tab. 1. taphonomic grades for daphnia caudal spines. excellent preservation fair preservation poor preservation stiff and straight spines with stiff and straight spines with slack spines. initial shape is no visible damage breakage and/or folding no longer sustained fig. 1. a) an example of a well-preserved caudal spine graded as excellent. b) a relatively well preserved, but visibly damaged caudal spine graded as fair. c) a poorly preserved caudal spine graded as poor. no n c om me rci al us e o nly varying preservation of subfossil daphnia longispina 145 diatom-inferred lake water ph (di-ph) the ph history of kivijärvi was quantitatively reconstructed using an independent modern diatom-water ph calibration data set consisting of 98 surface-sediment diatom assemblages from northern finland and corresponding ph measurements (for more details see seppä and weckström, 1999; väliranta et al., 2011). after testing different models (wa-inverse deshrinking, wa-classical deshrinking, partial-least-squares) the 1-component weighted average partial least squares (wa-pls) model provided the best performance with a coefficient of determination (r2) of 0.68 and a root mean square error of prediction (rmsep) of 0.31 ph units. this diatom-based quantitative ph model was then applied to the fossil diatom data analyzed from lake kivijärvi. methods and taxonomic literature used for diatom analyses are described in weckström et al. (1997). the quantitative diph is inferred only for the depths of 1-16 cm, as the original reason for constructing the di-ph model was to study the recent impact of the talvivaara mine on the lake kivijärvi water chemistry (unpublished data). results dating and sediment geochemistry according to the radiometric dating, the sediment depth of 2.5 cm records the fallout from chernobyl reactor accident (ad 1986) and the depth of 4.5 cm records the fallout maximum from atmospheric nuclear weapons testing (ad 1963, appleby et al., 1991). the dating provided a sedimentation rate of approximately 0.1 cm year–1. unsupported lead concentrations reach zero values at the depth of 10 cm, which corresponds to the beginning of the 20th century (1910±8). concentrations of k and mg increased after 1925 and reached maximum values in the late 1990s (fig. 2). fig. 2. results of taphonomical stratigraphy, geological characteristics, predation indices and diatom-inferred lake ph. eubosmina measurements are in µm (bar, average carapace length or mucro length; vertical line inside bar, standard deviation). note that eubosmina were not measured between depths 1-5 cm. chaoborus spp., number of chaoborus spp. claws counted in each sample per 100 daphnia spines. horizontal line indicates the statistically significant shift in the preservation pattern. no n c om me rci al us e o nly 146 j.j. leppänen and j. weckström daphnia remains many daphnia remains were detected throughout the core, except for the top 0-1 cm section, where they were extremely rare, likely in relation to the impact of the mine pollution. in the other sediment layers, postabdominal claws, carapace, ephippia, caudal spines and head shields were present in large numbers. only daphnia longispina o.f. müller -type postabdominal claws were detected. head shields showed various degrees of preservation including many very well preserved helmet-type head shields (fig. 3a), with one head shield still attached to the carapace (fig. 3b). there was a significant increasing trend towards present in the proportion of caudal spines graded as excellent (mann-kendall: z=3.88, s=103, p<0.001) and fair (mann-kendall: z=3.53, s=94, p<0.001), and a significant decreasing trend in the proportion of caudal spines graded as poor (mann-kendall: z=3.94, s=-105, p<0.001). zoning revealed two significant zones at depths 1-9 and 9-26 cm (fig. 2). predation assessment the length of the mucro and carapace of e. longispina exhibited no clear shifts within the core (fig. 2). measurements were terminated at the depth of 5 cm because of the appearance of bosmina longirostris o. f. müller in the sediment, as carapaces of e. longispina and b. longirostris are extremely difficult to be discriminated from each other. chaoborus mandibles were detected in very small numbers (0-2 mandibles per sample) throughout the core. di-ph the di-ph of the deepest core layers (which were likely deposited during the 18th century), was comparable to di-ph values around year 2000 (fig. 2). however, diph started to decrease since the late 19th century reaching the lowest values during the 1960s (fig. 2). di-ph remained at lower values between ca. 1920 and 1995, but later increased towards modern levels. discussion daphnia remains in lake kivijärvi usually only daphnia postabdomal claws and ephippia are found in lake sediments. sometimes, as reported by frey (1991), mancini et al. (1999), and sarmaja-korjonen (2007), different components of daphnia subfossils can stay relatively intact in the sediment for extended periods of time. also daphnia head shields have been detected in sediment samples (frey, 1991; mancini et al., 1999), but to our knowledge, helmet-type head shields of daphnia longispina –group species have not been reported from lake sediment samples before. in the sediments of lake kivijärvi sediments, the daphnia head shields appeared folded with fornices clearly visible, and fig. 3. a) daphnia helmet type headshields. scale bar: 100 µm. b) daphnia headshield attached to carapace. no n c om me rci al us e o nly varying preservation of subfossil daphnia longispina 147 a number of different structures were well preserved. however, the caudal spines showed the highest abundance. even though postabdominal claws are usually used in research on subfossil daphnia (korosi et al., 2011), caudal spines were used in the present work, as they could be classified more easily due to their simpler form, while the classification of claws in relation to physical condition would have been very difficult due to the higher variability in their characteristics and damage magnitude. in addition, caudal spines exhibited one additional character of degradation, namely the total loss of form (slack spines). the spines graded as poorly preserved were slack, suggesting that their structure had been altered. according to tang et al. (2009), cladoceran carcasses are rapidly and heavily colonized by chitin-degrading bacteria. loss of form may therefore refer to microbial or chemical degradation, rather than to a mechanical impact. in addition, the loss of form can be considered as linked to autochthonous in-sediment processes, because, if the loss of form were a result of sample preparation, it should have affected similar proportions of spines in all samples. in contrast, spines graded as fair good preserved were almost intact apart from small breakages likely resulting from sample treatment (e.g., sieving) or from predator or scavenger attacks. due to the uncertainty in pointing out whether the damage originates from laboratory treatment or from autochthonous processes, the indicator value of fair graded spines in this study is negligible. changes in daphnia preservation in lake kivijärvi according to the statistically significant core zonation, the greatest shift in daphnia preservation occurred at the depth of 9 cm (~1925) and was clearly reflected by changes in proportions of different grades. though not statistically significant, also the depth of 12 cm (~ pre 1900) clearly emerged as the level where early changes in daphnia preservation onset. the 9 cm shift is most pronounced for the excellent and poor grades, whereas the change regarding spines graded as fairly preserved is not as distinct. earlier studies suggest that daphnia preservation is controlled by temperature (szeroczyñska and zawisza, 2005) or water chemistry (sarmaja-korjonen, 2007). however, changes in hypolimnetic temperatures of lake kivijärvi during the 1920s are not likely, because a most notable warming occurred in n-finland only after the 1960s (mikkonen et al., 2014). moreover, hypolimnetic temperatures during stratification are not easily altered by relatively small scale increase in summer air temperatures (arvola et al., 2010). di-ph showed a slight decrease since ~1910, which may have been related to natural acidification i.e. continuous leaching of buffering mineral elements like calcium (ca), potassium (na), magnesium (mg), and sodium (k) (pennington, 1984). in fact, the proportion of naturally acidified lakes is higher in central and norther finland compared to s finland, where anthropogenic acidification is more common (meriläinen and huttunen, 1990). the decrease in di-ph since ca. 1910 corresponds relatively well to the change in daphnia preservation, but effects of ph on preservation changes is not straight forward. in fact, preservation was not affected when ph increased since ca. 1995, thus contradicting the response during slight acidification at the beginning of the 20th century. the lowest di-ph values around 1960 could be due to the increased impact of acid rain on small freshwaters as sulphur emission in finland increased substantially during the period of 1950-1970 (kauppi et al., 1990). however, this period of lower ph seemed not to affect the preservation of daphnia remains (fig. 2). the level of preservation of all subfossil material is usually related to accumulation rates, with good preservation (high taphonomic grade) reflecting fast accumulation and burial (brandt, 1989). since the core was only 210pb-dated, we lack dates previous to 1900 (i.e., below the depth of 10 cm). as a consequence, the possibility of low sedimentation rates and subsequent poor preservation of daphnia before ca. 1925 cannot be ruled out. possible reasons behind major changes in the sedimentation rate during the pre-industrial era should be principally related to hydrological perturbations, such as water level manipulation and channel building. however, according to the map by gylden (1848), shape and size of kivijärvi did not change since the mid-19th century, which excludes any large-scale water works during the early 20th century. cladoceran remains are directly affected by microbial degradation of chitin, which occurs both in the water and in the sediment (swiontek brzezinska et al., 2008). bacteria are the most important responsible of chitin decomposition in aquatic environments (aumen, 1980; gooday, 1990), but there are indications that particularly daphnia remains are attacked also by aquatic fungi (czezuga et al., 2002). in addition, chemical changes of the sediments (diagenesis) may alter the conditions within the sediment and the degree of ongoing degradation of chitinous subfossil remains (kidwell and flessa, 1995). forestry activities may affect lake water characteristics such as chemical oxygen demand (ahtiainen, 1992; rask et al., 1998), which in their turn may affect the fungal community (wurzbacher et al., 2010) and slow down the bacterial decomposition of chitin (köllner et al., 2012) at the sediment-water interface. however, documented forestry activities occurred well after the 1920s, while the recent increasing concentrations of erosion indicators (virkkanen and tikkanen, 1998) in the studied sediment core started only during the last decades of the 19th century. in addition to forestry actions, also changes in lake productivity could affect microbial community (naeher et al., 2012). however, according to the available maps and perno n c om me rci al us e o nly 148 j.j. leppänen and j. weckström sonal communication by s. peronius, no farmland existed within the catchment during the early 20th century. moreover, the first residents occupying the lake´s shoreline in the early 20th century were not farmers, but employed, thus the agricultural land use was minimal. this is also supported by the diatom data (not shown here), which did not show any indication of eutrophication as no taxa preferring elevated nutrient concentrations occurred during the last centuries. the sole known potentially relevant event that occurred in lake kivijärvi during the 1920s was the beginning of intense fishing activity, which resulted in large vendace (coregonus albula linneaus) catches since the 1920s (s. peronius, personal communication). planktivorus fish are known to target large-size zooplankton zaret, 1980), and vendace has been noted to prefer large cladocerans, such as eubosmina and daphnia, in swedish (hamrin, 1983) and finnish (viljanen, 1983) forest lakes. cladoceran remains are poorly digested during their passage through fish guts (sutela and huusko, 1993; riccardi, 2000) and daphnia remains have been proved to survive the passage through vendace and whitefish digestive apparatus (sutela and huusko, 2000). the poor preservation level of daphnia spines before the 1920s may therefore reflect a higher proportion of partly digested daphnia in the correspondent sediment samples. the intensive removal of fish may have also increased the abundance of invertebrate predators (milardi et al., 2016). elevated invertebrate predation has been noted to induce thickening and hardening of daphnia carapaces (rabus et al., 2013). this might be the reason for the appearance of hard and thick remains, which are more resistant to post-mortem degradation. however, the lack of variations of eubosmina size suggests that no high magnitude changes in invertebrate predation have occurred during the shift in daphnia preservation. moreover, the very low number of chaoborus mandibles in our samples (<2 per sample) allow no assumptions regarding changes of predation regime in lake kivijärvi, because according to quinlan and smol (2010) a minimum of 5 to 10 mandibles is needed to reliably assess chaoborus assemblages. high head helmet and a long spine are also considered as general defensive structures, which are grown by daphnia when they are subjected to invertebrate predation (laforsch and tollrian, 2004). the fact that only daphnia remains with long spine and high helmet were identified throughout the core studied, further supports the hypothesis that only little change in the intensity of invertebrate predation occurred at lake kivijärvi during the last ~250 years. the results presented here suggest the potential of daphnia caudal spine to be used as indicator of cladoceran degradation in sediments. however, more work is still necessary to further evaluate this potential. for example, type and degree of degradation of caudal spines caused by fish ingestion, should be experimentally assessed. similarly, the microbial degradation of daphnia remains should be tested in a controlled environment, in order to clarify the degradation process and the actual microbial community involved in daphnia degradation. conclusions this work provides first evidence for the potential use of caudal spine as indicator of daphnia degradation in sediments. historical information and sediment geochemistry suggest that no large-scale environmental or hydrological changes have affected the remote lake kivijärvi during the early decades of the 20th century, when major changes in daphnia preservation occurred. on the other hand, the historical data and the diatom-inferred ph profile indicate the increase in fishing activity after 1920 and the steady decrease in water ph from the beginning of 20th century till the 1960s as possible drivers of spine preservation of daphnia. though more work is clearly needed to experimentally verify the role of these factors in affecting daphnia preservation, these preliminary results shed new light on the issue of large cladoceran preservation in lake sediments. this aspect is still poorly explored, but might have the potential to improve the reliability of palaeolimnological reconstructions. acknowledgments this work was funded by tellervo and jussi walden foundation, soil protection and environmental protection technology (mutku) society and kainuu centre for economic development, transport and the environment. we are grateful to jussi leppänen for assistance in the field and to two anonymous reviewers for their valuable comments and constructive suggestions which greatly improved this manuscript. references ahtiainen m, 1992. the effects of forest clear-cutting and scarification on the water quality of small brooks. hydrobiologia 243/244:465-473. appleby pg, nolan pj, gifford dw, godfrey mj, oldfield f, anderson nj, battarbee rw, 1986. 210pb dating by low background gamma counting. hydrobiologia 141:21-27. appleby pg, richardson n, nolan pj, 1991. 241am dating of lake sediments. hydrobiologia 214:35-42. arvola l, george g, livingstone dm, järvinen m, blenckner t, dokulil mt, jennings e, aonghusa cn, noges p, noges t, weyhenmeyer ga, 2010. the impact of the changing climate on the thermal characteristics of lakes, p. 85-101. in: g. george (ed.), the impact of climate change on european no n c om me rci al us e o nly varying preservation of subfossil daphnia longispina 149 lakes, aquatic ecology series 4. springer science+business media b.v. aumen ng, 1980. microbial succession on a chitinous substrate in a woodland stream. microbial ecol. 6:317-327. bennett k, 1996. determination of the number of zones in a biostratigraphical sequence. new phytol. 132:155-170. birks hjb, gordon ad, 1985. numerical methods in quaternary pollen analysis. academic press, london: 317 pp. brandt ds, 1989. taphonomic grades as a classification for fossiliferous assemblages and implications for paleoecology. palaios 4:303-309. czeczuga b, kozłowska m, godlewska a, 2002. zoosporic aquatic fungi growing on dead specimens of 29 freshwater crustacean species. limnologica 32:180-193. deevey es, 1964. preliminary account of fossilization of zooplankton in rogers lake. int. ver. theor. angew. limnol. verh. 15:981-992. ebert d, 2005. ecology, and evolution of parasitism in daphnia. national library of medicine (us), national center for biotechnology information, bethesda: 110 pp. eggermont h, martens k, 2011. preface: cladocera crustaceans: sentinels of environmental change. hydrobiologia 676: 1-7. european union, 2000. directive 2000/60/ec of the european parliament and the council of 23 october 2000 on establishing a framework for community action in the field of water policy. eur. j. commun. l327:1-72. frey dg, 1986. cladocera analysis, p. 667-692. in: b.e. berglund (ed.), handbook of palaeoecology and palaeohydrology. j. wiley & sons ltd. frey dg, 1991. first subfossil records of daphnia headshields and shells (anomopoda, daphniidae) about 10 000 years old from northernmost greenland, plus alona guttata (chydoridae). j. paleolimnol. 6:193-197. gilbert ro, 1987. statistical methods for environmental pollution monitoring. van nostrand reinhold, new york: 336 pp. gooday gw, 1990. the ecology of chitin degradation. adv. microb. ecol. 11:387-430. gylden cw, 1863. [sydöstra delen af uleåborgs län med norra delen af kuopio län].[map in finnish]. sektionen d4. karta öfver finland. hammer ø, harper dat, ryan pd, 2001. past: paleontological statistics software package for education and data analysis. palaeontol. electron. 4:1-9. hamrin sf, 1983. the food preference of vendace (coregonus albula) in south swedish forest lakes including the predation effect on zooplankton populations. hydrobiologia 101:121-128. irannezhad m, marttila h, kløve b, 2014. long-term variations and trends in precipitation in finland. int. j. climatol. 34:3139-3153. johnsen gh, raddum gg, 1987. a morphological study of two populations of bosmina longispina exposed to different predation. j. plankton res. 9:297-304. kauppi s, mannio j, hellsten s, nysten t, jouttijärvi t, huttunen m, ekholm p, tuominen s, porvari p, karjalainen a, saraaho t, saukkoriipi j, maunula m, 2013. [arvio talvivaaran kaivoksen kipsisakka-altaan vuodon haitoista ja riskeistä vesiympäristölle. suomen ympäristökeskuksen raportteja 11/2013].[book in finnish]. suomen ympäristökeskus, helsinki: 90 pp. kerfoot wc, 1981. long-term replacement cycles in cladoceran communities: a history of predation. ecology 62:216-233. kersalo j, pirinen p, 2009. [the climate of finnish regions].[book in finnish]. finnish meteorological institute. yliopistopaino, helsinki: 185 pp. kidwell sm, flessa kw, 1995. the quality of fossil record: populations, species and communities. annu. rev. ecol. syst. 26:269-299. korhola a, 1999. distribution patterns of cladocera in subarctic fennoscandian lakes and their potential in environmental reconstruction. ecography 22:357-373. korosi jb, jeziorski a, smol jp, 2011. using morphological characters of subfossil daphniid postabdominal claws to improve taxonomic resolution within species complex. hydrobiologia 676:117-128. korosi jb, kurek j, smol jp, 2013. a review on utilizing bosmina size structure archieved in lake sediments to infer historic shifts in predation regimes. j. plankton. res. 35:444-460. köllner ke, carstens d, keller e, vazquez f, schubert cj, zeyer j, bürgmann h, 2012. bacterial chitin hydrolysis in two lakes with contrasting trophic statuses. appl. environ. microbiol. 78:695-704. laforsch c, tollrian r, 2004. inducible defenses in multipredator environments: cyclomorphosis in daphnia cucullata. ecology 85:2302-2311. lampert w, 2011. daphnia: development of a model organism in ecology and evolution. in: o. kinne (ed.), excellence in ecology, book 21. international ecology institute. kauppi p, anttila p, kenttämies k, 1990. introduction to this book, p. xiii-xviii. in: p. kauppi, p. anttila, k. kenttämies (eds.), acidification in finland. springer-verlag. lotter, af, juggins s, 1991. ploprof, tran and zone. programs for plotting, editing and zoning of pollen and diatom data. inqua commmission for the study of the holocene, working group on data handling methods, newsletter 6. mancini m, comoli p, margaritora fg, 1999. an unusual type of daphnia head shields from plankton and sediments of himalayan lakes. j. limnol. 58:29-32. meriläinen j, huttunen p, 1990. lake acidification in finland. phil. trans. r. soc. lond. b 327:423-425. mikkonen s, laine m, mäkelä hm, gregow h, tuomenvirta h, lahtinen m, laaksonen a, 2014. trends in the average temperature in finland, 1847-2013. stoch. env. res. risk. a. 29:1521-1529. milardi m, siitonen s, lappalainen j, liljendahl a, weckström j, 2016. the impact of trout introductions on macro-and micro-invertebrate communities of fishless boreal lakes. j. paleolimnol. 55:273-287. naeher s, smittenberg rh, gilli a, kirilova ep, lotter af, schubert cj, 2012. impact of recent lake eutrophication on microbial community changes as revealed by high resolution lipid biomarkers in rotsee (switzerland). org. geochem. 49:86-95. nykänen m, vakkilainen k, liukkonen m, kairesalo t, 2009. cladoceran remains in lake sediments: a comparison between plankton counts and sediment records. j. paleolimnol. 42:551-570. pennington w, 1984. long-term natural acidification of upland sites in cumbria: evidence from post-glacial lake sediments, p. 28-46. in: freshwater biological association (eds.), fiftyno n c om me rci al us e o nly 150 j.j. leppänen and j. weckström second annual report for the year ended 31st march 1984. freshwater biological association, ambleside. peters rh, de bernardi r, 1987. daphnia. mem. ist. ital. idrobiol. 45:1-502. quinlan r, smol jp, 2010. the extant chaoborus assemblage can be assessed using subfossil mandibles. freshwater biol. 55:2458-2467. rabus m, söllardl t, clausen-schaumann h, laforsch c, 2013. uncovering ultrastructural defences in daphnia magna – an interdisciplinary approach to assess the predator-induced fortification of the carapace. plos one 8:e67856. rask m, nyberg k, markkanen s-l, ojala a, 1998. forestry in catchments: effects on water quality, plankton, zoobenthos and fish in small lakes. boreal env. res. 3:75-86. renberg i, hansson h, 2008. the hth sediment corer. j. paleolimnol. 40:655-659. riccardi n, 2000. comparison of different stoichiometric methods for the estimation of proximate biochemical composition of crustacean zooplankton and some considerations on energy transfer to planktophagous fish. j. limnol. 59:179-185. sarmaja-korjonen k, 2002. multi-proxy data from kaksoislammi lake in finland: dramatic changes in the late holocene cladoceran assemblages. j. paleolimnol. 28:287-296. sarmaja-korjonen k, 2007. subfossil shell margins and caudal spines of daphnia in finnish lake sediments – is daphnia underrepresented in cladoceran analysis? studia quaternaria 24:61-64. schmidt r, wunsam s, brosch u, fott j, lami a, löffler h, marchetto a, müller mw, pražáková m, schwaighofer b, 1998. late and post-glacial history of meromictic längsee (austria), in respect to climate change and anthropogenic impact. aquat. sci. 60:56-88. seda j, petrusek a, 2011. daphnia as a model organism in limnology and aquatic ecology: introductory remarks. j. limnol. 70:337-344. seppä h, weckström j, 1999. holocene vegetational and limnological changes in the fennoscandian tree-line area as documented by pollen and diatom records from lake tsuolbmajavri, finland. ecoscience 6:621-635. smol jp, 1992. paleolimnology: an important tool for effective ecosystem management. j. aquat. ecosyst. health 1:49-58. sprules wg, carter jch, ramcharan cw, 1984. phenotypic associations in the bosminidae (cladocera): zoogeographic patterns. limnol. oceanogr. 29:161-169. sutela t, huusko a, 1994. digestion of zooplankton in the alimentary tract of vendace (coregonus albula) larvae. j. fish. biol. 44:591-596. sutela, t, huusko a, 2000. varying resistance of zooplankton prey to digestion: implications for quantifying larval fish diets. t. am. fish. soc. 192:545-551. sweetman jn, smol jp, 2006. reconstructing fish populations using chaoborus (diptera:chaoboridae) remains – a review. quat. sci. rev. 25:2013-2023. swiontek brzezinska m, lalke-porczyk e, donderski w, 2008. utilization of shrimp waste as a respiration substrate by planktonic and benthic microorganisms. pol. j. environ. stud. 17:273-282. szeroczyñska k, zawisza e, 2005. daphnia remains from the sediment of lake somaslampi (nw finnish lapland) and lake wigry (ne poland). studia quaternaria 22:55-57. szeroczyñska k, sarmaja-korjonen k, 2007. atlas of subfossil cladocera from central and northern europe. friends of the lower vistula society: 84 pp. tang kw, bickel sl, dziallas c, grossart h-p, 2009. microbial activities accompanying decomposition of cladoceran and copepod carcasses under different environmental conditions. aquat. microb. ecol. 57:89-100. viljanen m, 1983. food and food selection of cisco (coregonus albula l.) in a dysoligotrophic lake. hydrobiologia. 101: 129-138. virkkanen j, tikkanen m, 1998. the effects of forest ditching and water level changes on sediment quality in a small lake, perhonlampi, central finland. fennia 176:301-317. väliranta m, weckström j, siitonen s, seppä h, alkio j, juutinen s, tuittila e-s, 2011. holocene aquatic ecosystem change in the boreal vegetation zone of northern finland. j. paleolimnol. 45:339-352. weckström j, korhola a, blom t, 1997. the relationship between diatoms and water temperature in thirty subarctic fennoscandian lakes. arctic alpine res. 29:75-92. wurzbacher cm, barlocher f, grossart hp, 2010. fungi in lake ecosystems. aquat. microb. ecol. 59:125-149. zaret tm, 1980. predation and freshwater communities. yale university press, new haven: 185 pp.no n c om me rci al us e o nly layout 1 advances in oceanography and limnology, 2015; 6(1/2): 58-75 original article doi: 10.4081/aiol.2015.5470 introduction coastal zones are intensively populated by humans worldwide. today, approximately half of the world’s population lives within 200 kilometres from the oceans’ coastline, and this trend is likely to continue in next decades (creel, 2003). as a consequence, most anthropogenic activities are concentrated along the coast, resulting in multiple impacts on both terrestrial and marine environments. in the adriatic sea, the densely populated coast of the emilia romagna region hosts a plethora of human activities including, among others, fishing, tourism, shellfish farming, maritime transport and gas extraction. this area is also influenced by the po river, the most important freshwater input of the adriatic sea, and other secondary watercourses that contribute to significant inputs of nutrients, of both natural and human origin, to the sea. thus, the benthic domain along the emilia-romagna coast is subjected to a wide spectrum of stressors all likely affecting benthic communities. during the 80ies, the increased nutrient inputs from the po river caused acute alteration events ascribable to eutrophication, including red tides, anoxic bottom conditions and mass mortalities of benthic animals (crema et al., 1991 and references therein). nevertheless, in the last decade, the frequency of such events decreased likely due to the reduction of river nutrient loads (especially phosphates) and due to a sensible decrease in river runoff. consequently, in the northern adriatic sea, the general pattern seems to have switched from the eutrophic situation of the 80ies towards oligotrophication (mozetič et al., 2010). in the benthic domain, microphytobenthos (mpb), meiofauna and macrofauna are closely linked one to other through numerous interactions, which contribute to the complexity of the benthic food web (albertelli et al., 1999). mpb constitute the only autochthonous source of primary production on non-vegetated sediments of shallow marine areas (larson and sundbäck, 2008 and references therein). consequently, as principal oxygen producers, they allow the aerobic degradation of both autochthonous and settled organic matter (cibic et al., 2007 and references therein). moreover, mpb play a key role spatial distribution of microphytobenthos, meiofauna and macrofauna in the north-western adriatic sea: a synoptic study annalisa franzo,* tamara cibic, paola del negro, cinzia de vittor sezione oceanografia, istituto nazionale di oceanografia e geofisica sperimentale (ogs), via a. piccard 54, i-34151 trieste, italy abstract in view of the general paucity of integrated information on offshore benthic communities in the adriatic sea and given the vulnerability of this particular coastal system, microphytobenthos, meiofauna and macrofauna were synoptically investigated in front of the emilia-romagna coast (northern adriatic sea) in september 2010 and march 2011. as required by the marine strategy framework directive, which extends its action beyond the territorial waters (within 12 nmi) of the member states, our findings could help to fill the gap of knowledge on the environmental status in offshore areas since the study was carried out also at >12 nmi from the coastline. in fact, sediment samples for the analysis of the benthic communities were collected from a 10-point-station grid that covered an area of about 400 km2 with water depths ranging from 13 to 50 m. the variability of the sediment grain size and other chemical variables in the sediment suggests the presence of two distinct environmental contexts that enhance the proliferation of different benthic communities. at the inshore stations (depth <20 m) the higher sand percentages and the maxima of dissolved inorganic nitrogen concentrations indicated the presence of hydrodynamic perturbations and the influence of nutrient loads of terrestrial origin. inshore, both meioand macrofaunal communities were poorly structured and dominated by relatively more opportunistic taxa, such as nematodes and the bivalve corbula gibba. offshore stations (depth >20 m) had muddier sediments, which likely exerted a greater retention of sediment-bound organic matter. these conditions seemed to favour benthic deposit feeders like meioand macrofaunal annelids. surprisingly, a conspicuous microphytobenthic community, mainly represented by the diatom paralia sulcata, has been observed even at remarkable depths (~50 m) opening new questions regarding the role of these organisms in dim-light conditions. although the investigated benthic communities were, to some extent, moderately influenced by multiple stressors, the co-existence of sensitive taxa (such as meiofaunal copepods and the amphipod ampelisca spp.) with opportunistic ones (i.e., c. gibba) suggests that the benthic domain in the investigated area is not severely impacted. key words: microphytobenthos; meiofauna; macrofauna; adriatic sea; offshore benthos. received: july 2015. accepted: december 2015. no nco mm er cia l u se on ly 59 a. franzo et al. in the carbon cycle in littoral environments as producers of new organic matter that can enter either the benthic or the pelagic trophic webs (barranguet, 1997). mpb can also regulate nutrient fluxes by acting as a sort of nutrient filter: photosynthetically active microalgae, in fact, by the uptake of inorganic nutrients from the sediment pore water, can reduce nutrient fluxes towards the overlying water (sundbäck et al., 2004). meiofauna, including different trophic groups (i.e., deposit feeders, microalgae feeders, bacterivores and predators; balsamo et al., 2010) and being able to influence nutrient regeneration (bonaglia et al., 2014), represent a fundamental link between organic detritus and higher trophic levels, therefore playing a paramount ecological role within the surface sediments (balsamo et al., 2010 and references therein). in addition, some taxa, preying upon juvenile macrofauna, which are temporarily part of meiofauna, exert a bottleneck mechanism, controlling both the structure and the composition of the macrofaunal community (danovaro et al., 1995). in turn, meiofaunal organisms are intensively exposed to predation pressure by several small macrobenthic animals such as small shrimps, crabs, polychaetes, ophiuroids, and juvenile fishes (balsamo et al., 2010 and references therein). sediment bioturbation due to both meioand macrofauna enhances oxygen diffusion rates favouring aerobic degradation of organic matter by prokaryotes (zonneveld et al., 2010). overall, all size classes of benthic communities can be suitable for investigating and characterising the habitat where they live. they are simultaneously exposed to several environmental stressors that may modify each of them in either direct or indirect ways. directly, by altering their abundances and species composition and, indirectly, by inducing further changes at other trophic levels, e.g. modifying predator-prey relationships or altering primary production and respiration (semprucci et al., 2015). although microphytobenthos, meiofauna and macrofauna have different (but partially overlapping) roles in the functioning of sedimentary habitats and show different response times to anthropogenic pressures, they altogether represent an ideal tool for assessing the state of the benthic environment (semprucci et al., 2013 and references therein). along the emilia-romagna coastline, some synoptic studies have been already conducted integrating the analyses of meioand macrofauna with prokaryotes (semprucci et al., 2010) or foraminifera (frontalini et al., 2011; semprucci et al., 2013). on the other hand, to the best of our knowledge, mpb have been only rarely taken into consideration in the adriatic sea (totti, 2003; cibic and facca, 2010). the unique available study refers to strictly shallow communities (depth <15 m), so that, to date, there is a serious lack of knowledge on the offshore and deeper ones. there is an urgency to fill such gap in order to face the requirements of one of the main european umbrella regulations for waters systems, i.e. the marine strategy framework directive (msfd, 2008/56/ec, european commission, 2008). the msfd, in fact, extends its action to offshore marine environments, located beyond the territorial waters of the member states, i.e. at more than 12 nmi from the coastline, since even such areas could be affected by anthropogenic stressors. to provide new insights about offshore benthic communities in the adriatic sea, we conducted a study in front of the emilia-romagna coast (northern adriatic sea) in a wide area that partially protrudes beyond the italian territorial waters. structure and taxonomic composition of mpb (with a focus on diatoms), meiofauna and macrofauna were investigated synoptically and related to the characteristics of the sedimentary environment. more precisely, the aim of the present study was to answer the following questions: 1) what are the spatial distribution patterns of mpb, meiofauna and macrofauna along a water depth gradient? 2) is there an indication of possible disturbance factors affecting the benthic domain? methods study area the northern adriatic sea is a shallow sub-basin of the adriatic characterised by a morphologically complex coastline, which leads to variable hydrodynamic and sedimentary environments. the hydrology is influenced by several forcing factors like winds and river runoffs, and the general circulation is cyclonic. freshwater (mean flow rate of 1496 m3 s–1 in the period 1917-2008, cozzi and giani, 2011) is discharged into the basin from major rivers along the northern and north-western coasts. among them, the po river represents the major input of nutrientrich waters, accounting for about one third of the total riverine freshwater input in the adriatic. its plume, characterised by relatively fresh and mesotrophic waters, can spread over the entire northern sub-basin as a thin surface layer in summer while it is reduced and flows directly southward along the italian coast in winter (poulain et al., 2001 and references therein). the sedimentation pattern matches the hydrodynamic circulation (tomadin, 2000; ravaioli et al., 2003). it consists of a narrow strip of recent sands along the coast followed by a broad belt of muddy sediments. going offshore, there is a transitional muddy zone characterised by a gradual increase of sand fraction, which becomes dominant in a wide open sandy shelf area with little or no recent sedimentation (known as relict sands) (colantoni et al., 1985). although the northern adriatic shelf is a relatively low-energy environment with small tidal ranges and wave heights, after flood events, ephemeral deposits of sediments are generated and subsequently remobilized no nco mm er cia l u se on ly 60synoptic study of three benthic communities by waves to form dense flows (traykovski et al., 2007). the sediments travel southwards in a series of wind-induced resuspension events promoted by the bora and scirocco winds (fain et al., 2007). the main sources of biogenic elements are autochthonous (plankton) or allochthonous (atmospheric inputs and organic matter derived from soil and transported by rivers). the investigated area is located in front of the emilia-romagna coast, an area impacted by several anthropogenic activities, such as the presence of offshore gas platforms (manoukian et al., 2010), sand extraction for beach replenishment (simonini et al., 2007) and the dumping of harbour-dredged materials (simonini et al., 2005b). sampling the study was carried out in a coastal area between rimini and pesaro cities. to characterise both the spatial and temporal variability of chemical, biological and physical parameters in the sediment, a 10-point-station grid, covering an area of about 400 km2, was sampled in september 2010 and march 2011 (fig. 1). st. c01 was located 5.5 km (2.7 nmi) from the coast, while st. b02, c02 and d02 were, on average, 14.4 km (7.4 nmi) offshore. st. b03, c03, d03 were 20.8 km (10.7 nmi) from the coast and st. b04, c04, d04 were located, on average, 27.0 km (13.8 nmi) offshore. the depth ranged from ~14 m to ~50 m at st. c01 and st. d04, respectively (tab. 1). in the present study, the depth of 20 m was considered as the threshold between coastal and offshore stations. therefore, st. c01, b02, c02 and d02 are referred as inshore stations throughout the paper whereas st. b03, c03, d03, b04, c04 and d04 as offshore stations. in the present study, september and march were chosen as sampling periods because representative of summer and winter conditions for the considered benthic communities in the northern adriatic sea (cibic et al., 2012; n’siala et al., 2008 and references therein). for logistical reasons, sediment samples for chemical and biological data were collected within a 15-day period during the september survey and within a 20-day period in march. for grain size, chemical variables and pigments, fig. 1. location of the study area and structure of the 10-point-station grid. image drawn by means of odv software (schlitzer, 2015). no nco mm er cia l u se on ly 61 a. franzo et al. virtually undisturbed sediment cores were collected by a kc haps bottom corer (kc-denmark, silkeborg, denmark) using plexiglas sample tubes (16 cm i.d; 30 cm length). sediment cores were partially extruded in a n2filled chamber and the oxic layer (0-1 cm) was collected and homogenised. for dissolved nutrients analysis, pore waters were extracted from surface sediments by centrifugation (3500 rpm, 40 min hereaous omnifuge 2000 rs) at in situ temperature, subsequently recovered under a n2 atmosphere and filtered through 0.45 µm pore cellulose acetate filters. for logistical difficulties, the chemical data from st. c02 are not available in march. for mpb and meiofauna, sediment was collected using a box corer (25 cm height; 17x10 cm of sampling area). sediment cores were subsampled for meiofauna using cut-off plastic syringes (2.7 cm i.d., length 11.4 cm) in three replicates at each station, and the remaining oxic layer was collected for microphytobenthic analyses. macrofauna were sampled in three replicates per station using a van veen grab with a 0.1 m2 sampling area. depth profiles of physical characteristics of the water column were recorded by a seabird 19 plusseacat probe. sediment grain size, chemical parameters and pigments for grain size analysis, aliquots of sediment (10-15 g) were collected at all stations. small pebbles and shell fragments were separated from sand and mud fractions by sieving at 2 mm. each sample was disaggregated and the organic matter oxidized using 10% hydrogen peroxide (carlo erba, rodano, italy) at 60°c for 24 h. the analyses were performed using a malvern mastersizer 2000 equipped with hydro 2000s (malvern instruments ltd., malvern, uk). data are expressed as percentages of sand, silt and clay. for total organic carbon (toc) and total nitrogen (tn) analyses, sediment was freeze-dried and grounded using a mortar, then a pestle and the fraction >250µm was separated from the rest of the sample. triplicate subsamples of about 8-12 mg were weighed directly in a capsule (5x9 mm) on a microultrabalance perkin elmer mod. ad4 (accuracy of 0.1 µg). for toc and tn analyses, silver capsules and tin capsules were used, respectively. before the toc determination, subsamples were treated with increasing concentrations of hcl (0.1n and 1n) to remove carbonates (nieuwenhuize et al., 1994). carbon and nitrogen content were determined using a chno-s elemental analyser mod. ecs 4010 (costech international spa, pioltello, italy) according to pella and colombo (1973) and sharp (1974). capsule blanks were also analysed. the precision of the analysis was about 3%. the c and n contents in the samples are expressed as mg c gdry mass–1and mg n gdry mass–1. dissolved inorganic nutrient concentrations (nitrite, ta b. 1 .t he p er ce nt ag e of m ud a nd th e ch em ic al p ar am et er s m ea su re d in th e po re w at er o f t he s ed im en t s ur fa ce la ye r ( 01 cm ) i n bo th p er io ds o f t he s tu dy . se pt em be r 20 10 m ar ch 2 01 1 d ep th m ud t o c t n d in p -p o 4 si o 32 – c hl a p ha eo m ud t o c t n d in p -p o 4 si o 32 – c hl a p ha eo st . m % m g c g –1 m g n g –1 µ m µ m µ m µg g –1 µ g g– 1 % m g c g –1 m g n g –1 µ m µm µm µ g g– 1 µ g g– 1 b 02 2 1. 2 8 7. 68 1 1. 36 1. 21 4 2. 69 1. 34 81 .7 2 2 .6 2 1 6. 68 9 0. 29 7 .2 8 0 .8 5 3 8. 70 0 .9 6 14 0. 02 0. 99 9 .4 1 b 03 3 1. 0 9 7. 04 1 0. 29 1. 26 4 5. 72 1. 56 1 36 .3 0 1 .0 7 1 9. 85 9 8. 29 10 .0 8 1 .3 5 1 9. 98 1 .0 2 12 4. 71 1. 41 1 6. 27 b 04 4 1. 7 9 8. 35 1 0. 36 1. 43 5 1. 43 1. 35 1 20 .3 7 0 .9 9 2 3. 24 9 8. 47 12 .0 6 1 .7 0 1 8. 79 1 .2 7 14 2. 04 2. 61 1 9. 75 c 01 1 3. 5 5 5. 16 1. 71 0 .1 4 9 8. 74 1. 21 51 .8 8 0 .2 9 3 .6 8 6 8. 96 4 .7 5 0 .6 1 3 1. 48 0 .6 2 9 .7 3 1 .4 7 8 .5 1 c 02 1 9. 0 7 4. 13 3. 83 0 .6 2 10 7. 97 0 .8 0 9 5. 09 1. 24 7 .9 5 n a n a n a n a n a n a n a n a c 03 3 4. 0 9 7. 49 1 3. 20 1. 58 7 4. 13 1. 59 1 48 .1 1 0 .6 6 1 4. 48 9 8. 01 10 .3 9 1 .2 3 1 4. 22 0 .5 7 6 3. 21 0 .5 3 1 0. 30 c 04 4 6. 4 8 8. 95 1 0. 09 1. 36 3 4. 42 1. 56 98 .3 0 0 .5 7 1 4. 12 9 4. 10 9 .1 0 1 .3 4 1 0. 21 1 .3 1 7 7. 40 1 .2 4 1 1. 47 d 02 1 9. 4 7 1. 29 4. 99 0 .6 0 9 4. 96 0. 83 92 .1 4 2 .1 8 1 5. 20 5 9. 73 3 .7 8 0 .5 1 5 .8 4 0 .6 5 7 .1 6 1 .4 8 8 .5 7 d 03 3 6. 0 9 7. 92 8. 05 0 .9 8 3 9. 90 1. 55 1 07 .0 0 0 .6 5 1 5. 66 9 7. 36 9 .4 2 1 .2 0 3 3. 76 1 .2 2 1 3. 90 0 .7 4 1 1. 35 d 04 5 0. 5 8 5. 03 9. 69 1 .0 7 2 4. 89 1. 63 1 06 .2 4 0 .6 1 1 7. 13 7 9. 03 8 .0 7 1 .2 0 2 0. 80 1 .5 2 10 2. 80 0. 93 9 .7 5 t o c , t ot al o rg an ic c ar bo n; t n , t ot al n itr og en ; d in , s um o f n h 4+ , n o 2– an d n o 3– co nc en tr at io ns ; p ha eo , p ha eo pi gm en ts ; n a, n ot a va ila bl e. no nco mm er cia l u se on ly 62synoptic study of three benthic communities no2–; nitrate, no3–; ammonium, nh4+; phosphate, po43–; silicate, sio32–) were analysed in the extracted pore water by a flow-injection spectrophotometric method (grasshoff et al., 1983). analyses were performed at room temperature on a five-channel bran+luebbeautoanalyzer 3 continuous flow analyzer (bran+luebbe, norderstedt, germany), using standard procedures. performance of the system was checked before and after sample analysis by running replicates of internal standards. the precision of the analysis was 3%. pigments were extracted overnight (4°c, 90% acetone) from 0.7-0.9 g of wet sediment. after centrifugation (2500 rpm, 20 min), the extract was analysed spectrofluorometrically following the procedures described by lorenzen and jeffrey (1980). the measurements were performed before (chlorophyll a, chla, µg g–1) and after (phaeopigments, phaeo, µg g–1) acidification with 2 drops of 1 n hcl, using a jasco fp 6500 spectrofluorometer (450 nm excitement and 665 nm emission wavelengths). calibration was made with pure sigma chl a standards. microphytobenthos for mpb, aliquots of 2 cm3 of surface sediment were withdrawn using a syringe and directly fixed with 10 ml of formaldehyde (4% concentration) buffered solution camg(co3)2 (carlo erba), in pre-filtered bottom seawater (0.2 µm filters) (schleicher & schuell, dassel, germany). after manual stirring, 20 µl aliquots of the sediment suspension were drawn off from the slurries and placed in a counting chamber (thalassia, trieste, italy). only cells containing pigments and not empty frustules were counted under a leitz inverted light microscope (leica microsystems ag, wetzlar, germany) (utermöhl, 1958) at 320x final magnification. the abundance was expressed as cells cm–3. the microalgal taxonomy was based on round et al. (1992), cibic and facca (2010) and the algaebase website (http://www.algaebase.org/). the qualitative identification of microphytobenthic assemblages was carried out to the genus and, when possible, to the species level using floras of van heurck (1899), hendey (1976), germain (1981), dexing et al. (1985), ricard (1987), tomas (1997) and witkowski (2000). meiofauna for the study of the meiofauna, three replicate cores were taken down to the depth of 10 cm. the sediment was fixed with 4% neutralized formaldehyde and stained with rose bengal 1‰ (vwr prolabo). sediment samples were sieved through 1000-38 µm mesh nets. the extraction of meiofaunal organisms from the sediment was performed by three times centrifugation with ludox am 30 (sigmaaldrich, st. louis, mo, usa) (density 1.18 g cm–3) as described by danovaro et al. (2004). the meiofauna were counted and identified under a stereomicroscope with a final magnification of 40 or 80x at the major taxonomic level according to higgins and thiel (1988); the abundance of the main groups was expressed as number of individuals per 10 cm2 (higgins and thiel, 1988). macrofauna sediment was sieved through 1000 µm mesh to retain the fraction of macrozoobenthic organisms and immediately frozen at -20°c on board (castelli et al., 2004). after defrosting, organisms were separated from sediment by tweezers, and firstly assigned to their taxon (i.e., molluscs, annelids, amphipods, etc.) and preserved in a borax buffered solution of 4% formaldehyde (carlo erba reagents). all the animals were identified to the lowest possible taxonomical level using a stereomicroscope at 740x final magnification and counted. for the identification of organisms, the taxonomical keys listed in morri et al. (2004) were used. the abundance was expressed as number of individuals per m2. statistical analyses diversity indexes d (richness of taxa; margalef, 1986), h’ (shannon and weaver, 1949) and λ (dominance; simpson, 1949) were calculated for benthic diatom and macrofaunal communities. to visualize differences among groups of stations (i.e., inshore vs offshore) cluster analyses based on square-root transformed data and bray-curtis similarity matrixes with a complete linkage were performed separately for each benthic group (i.e., mpb, meioand macrofauna), by pooling together data obtained in september and march (clarke and warwick, 2001). to assess whether the two groups of stations identified from the cluster analysis were statistically different among each other, anosim tests were performed considering such groups as the discriminating factor. further, variations across sampling periods (september vs march) and water depths (<20 m vs >20 m) were investigated using two-way crossed anosim tests applied on benthic diatoms, meiofauna and macrofauna datasets, separately. the same tests were applied also on normalized environmental data (sand, toc, tn, po43–, dissolved inorganic nitrogen din as sum of all inorganic nitrogen forms, chl a, phaeopigments) using euclidean distance matrixes. the outputs of anosim were interpreted according to clarke and gorely (2006): r<0.25=no differences among stations/sampling periods for the considered biotic/abiotic variables; r>0.5=indicated some overlapping but the stations/sampling periods differed for a certain degree; r>0.75=the stations/sampling periods were statistically different for the considered biotic/abiotic variables. when the two-way crossed anosim highlighted the presence of statistically differno nco mm er cia l u se on ly 63 a. franzo et al. ent benthic communities according to the tested factors, a similarity percentage (simper) analysis was carried out in order to detect which taxa were mainly responsible for the variations observed. a percentage dissimilarity cutoff of 50% was applied. diversity indexes, cluster, anosim and simper analyses were carried out using the routines included in the primer software ver. 5 (primer-e ldt, plymouth, uk). principal component analysis (pca) was carried out on the whole environmental and faunal dataset in order to visualise the trends of the main abiotic variables. after normalisation, the absolute abundances of benthic diatoms, meioand macrofauna were also projected on the factor plane as additional variables without contributing to the results of the analysis. this can provide an insight into the possible influence of the environmental variables upon each benthic group (statistica 7). to highlight interactions between biotic and abiotic variables a spearman’s rank correlation analysis (r) was performed using statistica 7 (p<0.05; p<0.01; p<0.001). results environmental variables in september, bottom temperatures varied from 11.9 °c (st. c04) to 23.4 °c (st. c01), with higher values at the shallower stations and colder conditions at the deeper ones. by contrast, in march, bottom temperatures were quite homogeneous (~10°c) within the entire 10-pointstation grid. at all stations, the uppermost sediment layer was mainly characterised by the mud fraction (as the sum of silt and clay percentages) during both study periods (tab. 1). in september, a lower percentage of mud was observed inshore, especially at st. c01, where the minimum was reached (55.2%). at the offshore stations, muddier sediments were found, particularly at st. b04, where the percentage of mud exceeded 98%. in march, overall similar grain-size features were observed with an increase of mud percentage from the shallower stations towards the open sea. surface mud fractions ranged from a minimum of 59.7% at st. d02 to a maximum of 98.5% at st. b04. according to shepard’s (1954) classification, the sediment was clayey silt at all stations, with the exception of st. c01 and st. d02 where it was sandy silt in both sampling periods. toc content varied between 1.71 mg c g–1 (st. c01) and 13.20 mg c g–1 (st. c03) in september and between 3.78 mg c g–1 (st. d02) and 12.06 mg c g–1 (st. b04) in march. tn content ranged from 0.14 mg n g–1 (st. c01) to 1.58 mg n g–1 (st. c03) in september and from 0.51 mg n g–1 (st. d02) to 1.70 mg n g–1 (st. b04) in march. coastal stations had lower toc and tn contents than the offshore ones, and similar values occurred in the two study periods. din concentrations measured in the pore water of the top sediment layer varied from 24.89 µm (st. d04) to 107.97 µm (st. c02) in september, and from 5.84 µm (st. d02) to 38.70 µm (st. b02) in march. din concentrations generally decreased from the inshore stations towards the open sea during both study periods. pore water po43– concentrations ranged from 0.80 µm (st. c02) to 1.63 µm (st. d04) in september, and from 0.57 µm (st. c03) to 1.52 µm (st. d04) in march. unlike din values, phosphate concentrations increased from the coast towards the open sea during both periods. the lowest sio32– concentrations were observed inshore (51.88 µm at st. c01 in september; 7.16 µm at st. d02 in march), whereas the highest values occurred at the offshore stations (148.11 µm at st. c03 in september; 142.04 µm at st. b04 in march). in september, chl a values ranged from 0.29 µg g–1 (st. c01) to 2.62 µg g–1 (st. b02). with the exception of st. c01, the shallower stations had the highest chl a sedimentary contents, whereas lower values were observed offshore. phaeopigments concentrations increased from the coast towards the open sea, with the lowest value observed at st. c01 (3.68 µg g–1) and the highest at st. b04 (23.24 µg g–1). in march chl a varied between 0.53 µg g–1 (st. c03) and 2.61 µg g–1 (st. b04), whereas phaeopigments between 8.51 µg g–1 (st. c01) and 19.75 µg g–1 (st. b04) (tab. 1). microphytobenthos in september, microphytobenthic abundance ranged from 9900±424 cells cm–3 at st. c01 to 86400±16971 cells cm–3 at st. d02 whereas in march the lowest value was recorded at st. b02 (18900±5515 cells cm–3), and the highest at st. c01 (111600±23759 cells cm–3). during both sampling periods, total mpb abundance was not correlated with water depth. the microphytobenthic community was mainly composed of diatoms (fig. 2 a,b). in september, bacillariophyta represented 45.5% of the total microalgal abundance at st. c01, whereas their ra reached 96.2% at st. d02 (fig. 2a). undetermined phytoflagellates and resting cells (spores and cysts) were found at all stations, whereas chlorophyceae were observed only at st. d03 and cyanophyceae only at st. c03 and d03. in march, diatoms relative abundance ranged from 66.7% at st. b02 to 94.6% at st. c01. relatively high percentages of phytoflagellates were observed at st. b04 (11.4%) and d03 (10.0%). undetermined spores were more abundant at st. b02, c02 and d03 than at all the other stations (fig. 2b, in legend as undet. spores). since, besides bacillariophyta, the specimens belonging to other groups remained mostly undetermined at higher taxonomic level, further analyses were focused only on the benthic diatom community. pooling the data together from all stations, 17 diatom genera were observed in september and 16 in march. in september, paralia was the dominant genus, followed by nitzschia, navicula and gyrosigma. in march, the no nco mm er cia l u se on ly 64synoptic study of three benthic communities planktonic genus skeletonema was very abundant at st. c01 and, overall, its abundance was higher at stations closest to the coast. paralia was the second most abundant genus, followed by navicula and diploneis (tab. 2). paralia sulcata was the most abundant diatom species in september, when reached a maximum of 33600 cells cm–3 at st. c04. fig. 2. relative abundance of microphytobenthos (a, b), meiofauna (c, d) and macrofauna (e, f) at each station along the 10-pointstation grid in september 2010 (left side) and march 2011 (right side). for meiofauna, taxa representing <1% of the total abundance are included in the group others and a cut-off at 50% on y-axis was applied. macrofauna taxa representing <10% of the total abundance at all stations are included in the group others. no nco mm er cia l u se on ly 65 a. franzo et al. rhopalodia gibba was found only at st. d02, c02 and b02, specifically in those stations where the depth ranged from 18.8 to 20.8 m. gyrosigma acuminatum and g. spencerii reached higher abundance at stations closest to the coast. in march, the planktonic species skeletonema costatum and the tychopelagic paralia sulcata reached the highest abundance followed by the benthic species diploneis bombus. the diatom community was more diverse in september than in march (tab. 4). in september, higher d and h’ values were observed at st. d02 (d=2.207; h’=2.782). the dominance of paralia sulcata at st. c04 (87.5% of the total diatom abundance) explained the highest λ observed at this station (0.770). in march, the highest diversity was recorded at st. b02 (h’=2.070) and the highest richness at st. d04 (d=1.289). the highest λ (0.891) was obtained at st. c01 due to the dominance of skeletonema costatum, which in this station accounted for 94.3% of the whole diatom community. the complete list of diatom species is reported in supplementary tab. 1. meiofauna in september, meiofauna ranged from 269.5±141.3 to 951.2±464.8 ind. 10 cm–2 at st. c03 and st. d02, respectively. in march, values varied from 376.8±113.6 ind. 10 cm–2 at st. b02 to 1030.9±70.0 ind. 10 cm–2 at st. b03. during both sampling periods, total abundance did not vary among inshore and offshore stations. a total of 16 major groups were identified, with nematodes being dominant in both study periods and at all stations (fig. 2 c,d). in september, these organisms represented 74.6% of the total meiofaunal abundance at st. d04 and up to 92.4% at st. d02, whereas in march their relative abundance varied from 69.6% (st. d03) to 96.0% (st. c01). copepods (adults and juvenile stages) represented the second most abundant taxon during both periods of the study and were more abundant offshore than inshore. their relative abundances ranged, in fact, from 5.3% (st. d02) to 19.8% (st. d04) in september and from 1.1% (st. c01) to 24.1% (st. d03) in march. the other main groups represented minor fractions of the whole community: polychaeta and kinorhyncha were observed only at stations with a depth >20m during both study periods. their relative abundances, in fact, did not exceed 1.6% (st. d03, september) and 2.2% (st. d03, march) of the total assemblage, respectively. undetermined organisms (fig. 2 c,d; in legend as incertae sedis) were observed at all stations in percentages generally <3% both in september and march. during both sampling periods meiofauna resulted more diverse at the offshore stations than at the inshore ones, with the latter more strongly dominated by nematodes. the number of taxa varied, in fact, from 4 (st. c01) to 11 (st. b04) in september. in march, the lowest number of taxa (4) was recorded at st. c01, and the highest (9) was observed at different depths (st. c03, st. c04 and st. d02). macrofauna shallower stations (b02, c01, c02 and d02) were characterised by higher macrofaunal abundance than the deeper ones during both sampling periods. values ranged from 76.7±30.6 ind. m–2 (st. b04) to 5166.7±1361.2 ind. tab. 2. abundance, expressed as cells cm–3, of the first ten dominant diatom species during the two periods of the study. b02 b03 b04 c01 c02 c03 c04 d02 d03 d04 september 2010 paralia sulcata 1800 2400 9900 300 0 2100 33600 0 7500 9000 rhopalodia gibba 900 0 0 0 4200 0 0 10800 0 0 nitzschia fasciculata 2700 0 0 0 600 0 0 10200 0 0 nitzschia lorenziana 2400 600 0 0 900 600 0 3900 300 0 navicula directa 1800 0 0 0 2700 0 0 2400 0 0 bacillaria paxillifera 0 0 0 0 0 0 0 6300 0 0 gyrosigma spencerii 3300 0 0 0 3000 0 0 0 0 0 gyrosigma acuminatum 1200 600 300 600 0 2100 0 0 300 300 pleurosigma aestuarii 0 0 0 0 2700 0 0 2400 0 0 nitzschia sigma 0 0 0 0 900 0 0 3000 0 0 march 2011 skeletonema costatum 4800 1800 0 99600 8700 0 3000 20700 300 3000 paralia sulcata 0 8100 16200 0 1500 18000 16200 0 7500 39300 diploneis bombus 300 2100 0 0 600 600 1200 0 1800 1500 gyrosigma acuminatum 300 0 900 0 0 600 600 0 300 1200 gyrosigma spencerii 900 300 0 300 300 0 0 600 0 0 ceratoneis closterium 0 0 0 900 0 0 0 900 0 300 rhopalodia gibba 300 0 0 300 0 0 0 1500 0 0 nitzschia tryblionella 600 0 600 0 300 0 0 0 300 0 entomoneis alata 0 0 0 300 0 900 0 0 0 0 thalassionema nitzschioides 0 0 0 0 0 1200 0 0 0 0 no nco mm er cia l u se on ly 66synoptic study of three benthic communities m–2 (st. c01) and from 123.3±110.2 ind. m–2(st. c04) to 1080.0±475.7 ind. m–2 (st. c02) in september and march, respectively. overall, the abundance in september was generally higher than in march, particularly at the shallower stations. macrofaunal taxa representing <10% of the total abundance (e.g., anthozoans, copepods, echinoderms, priapulids, sipunculids) were pooled in others (fig. 2 e,f). in september, bivalves and amphipods were more abundant at the inshore stations than at the offshore ones, representing 65.6% at st. d02 and 31.4% at st. b02, respectively while being completely absent at st. b04, c04 and d04 (bivalves) and at st. c04 and d04 (amphipods). in both sampling periods, polychaetes gradually became the dominant group with increasing water depth. the relative abundance varied, in fact, from 7.7% (st. c02) to 87.5% (st. d04) and from 9.3% (st. c02) to 65.1% (st. d04) in september and march, respectively. priapulids were observed only offshore. in march, the decreasing percentage of bivalves from the coast towards the open sea was not clear as that in september. both in september and march the inshore stations were dominated by the bivalve corbula gibba, which reached the highest abundance at the shallowest st. c01 (2540.0 ind. m–2) in september (tab. 3). the offshore stations were mainly characterised by the polychaete sternaspis scutata during both sampling periods (up to 83.3 ind. m–2 at st. b03 and c03 in march). in september also the second dominant species, the amphipod ampelisca spinipes, showed the highest abundance inshore (up to 623.3 ind. m–2 at st. b02). in march, the inshore stations were also characterised by the presence of the bivalve tellimya ferruginosa and the gastropod turritella communis with the highest values observed at st. d02 (313.3 ind. m–2) and c02 (200.0 ind. m–2), respectively, and the offshore ones by the gastropod hydrobia acuta (86.7 ind. m–2 at st. b03). in september, the highest diversity was observed at st. d02 (h’=2.520), which also had the second highest richness (d=6.561). the lowest values of both h’ (1.636) and d (2.190) were observed at st. d04 in correspondence with the highest dominance (λ=0.360). in march, the maximum diversity (h’=2.866) and the second lowest dominance value (λ=0.105) were observed at st. c01. the lowest richness (d=2.492) and diversity (h’=2.055) were observed at st. c04 in correspondence with the maximum dominance (λ=0.186) (tab. 4). the complete list of the observed macrofaunal organisms is reported in supplementary tab. 2. statistical analyses overall, the cluster analysis performed on the datasets of each benthic community separately showed the presence of distinct assemblages according to the different depth. on the contrary, since the subclusters comprised indiscriminately samples of september and march, the variability due to the different period seemed less important (fig. 3). the results of the anosim test indicated that, for all the three benthic communities, the groups identified by the cluster analysis were statistically different among each other (benthic diatoms, r=0.818; meiofauna, r=0.706; macrofauna, r=0.922). focusing on benthic diatoms, two clusters were tab. 3. abundances, expressed as ind. m–2, of the first ten dominant macrobenthic species during both periods of the study. b02 b03 b04 c01 c02 c03 c04 d02 d03 d04 september 2010 corbula gibba 523.3 0.0 0.0 2540.0 970.0 0.0 0.0 626.7 0.0 0.0 ampelisca spinipes 623.3 0.0 0.0 516.7 376.7 0.0 0.0 83.3 0.0 0.0 apseudopsis latreillei 50.0 0.0 0.0 743.3 30.0 3.3 0.0 0.0 0.0 0.0 tellimya ferruginosa 190.0 76.7 0.0 40.0 26.7 30.0 0.0 166.7 6.7 0.0 nucula sulcata 53.3 0.0 0.0 163.3 70.0 0.0 0.0 56.7 0.0 0.0 turritella communis 200.0 0.0 0.0 0.0 83.3 3.3 0.0 43.3 0.0 0.0 sternaspis scutata 43.3 43.3 10.0 0.0 20.0 80.0 16.7 10.0 40.0 20.0 amphictene auricoma 0.0 0.0 0.0 216.7 20.0 0.0 0.0 6.7 0.0 3.3 lumbrineris adriatica 63.3 3.3 0.0 136.7 16.7 0.0 0.0 13.3 0.0 3.3 hydrobia acuta 73.3 70.0 3.3 43.3 3.3 16.7 6.7 3.3 3.3 3.3 march 2011 corbula gibba 190.0 0.0 0.0 10.0 310.0 0.0 0.0 96.7 0.0 0.0 tellimya ferruginosa 0.0 53.3 3.3 40.0 40.0 30.0 0.0 313.3 16.7 6.7 turritella communis 56.7 0.0 6.7 0.0 200.0 10.0 0.0 110.0 6.7 3.3 hydrobia acuta 13.3 86.7 66.7 23.3 20.0 40.0 40.0 20.0 60.0 16.7 sternaspis scutata 66.7 83.3 26.7 0.0 3.3 83.3 30.0 13.3 63.3 10.0 ampelisca spinipes 43.3 13.3 0.0 190.0 110.0 0.0 0.0 3.3 3.3 0.0 nucula sulcata 23.3 0.0 0.0 0.0 63.3 0.0 0.0 113.3 0.0 0.0 caulleriella multibranchis 0.0 20.0 6.7 3.3 10.0 0.0 3.3 0.0 53.3 60.0 oestergrenia thomsonii 3.3 53.3 6.7 0.0 3.3 20.0 3.3 33.3 10.0 6.7 lumbrineris adriatica 3.3 6.7 0.0 63.3 20.0 0.0 0.0 23.3 3.3 3.3 no nco mm er cia l u se on ly 67 a. franzo et al. fig. 3. cluster analysis based on the abundances of benthic diatoms (a), meiofauna (b) and macrofauna (c) exploring for variations among stations and sampling periods. s, september 2010; m, march 2011. in red circles the outliers. no nco mm er cia l u se on ly 68synoptic study of three benthic communities discriminated by a depth threshold of 20 m: a cluster including all station at >20 m (plus st. c01 in september), and a cluster including all stations at <20 m depth (fig. 3a). the cluster analysis carried out for meiofauna grouped again stations in two clusters. though the discrimination according to water depth was less clear than the one observed for diatoms, a group included the majority of offshore stations (depth >20 m) and the other included the majority of inshore ones (depth <20 m) (fig. 3b). the cluster analysis carried out for macrofauna grouped the stations in two groups discriminated again by the threshold depth of 20 m (fig. 3c). following the results of the cluster analysis, two-way crossed anosim tests were carried out to test for differences in the composition of diatom, meiofauna and macrofauna communities, using sampling periods and water depth (i.e., >20 m and <20m depth) as discriminating factors. the outputs on the diatom community (tab. 5) revealed water depth as the unique factor able to explain significant community variations, while the sampling period seemed to not exert any kind of effect. the simper analysis revealed that the dissimilarity between inshore and offshore microalgal assemblages (depth <20 m vs depth >20 m) was mainly due to the presence of paralia sulcata (contrib.%=7.05) and skeletonema costatum (contrib.%=5.67): the first was more abundant offshore while the second species proliferated at shallower depth (tab. 6). meiofaunal communities did not differ significantly between either sampling times or water depth, whereas macrofauna community composition varied significantly among both sampling periods and water depth (tab. 5). the simper analysis on macrofauna showed that the main differences among sampling periods (irrespectively of sampling station) were due to the higher abundances of the dominant species in september (tab. 7). on the other hand, the dissimilarity between inshore and offshore communities (irrespectively of the sampling period) were mainly due to corbula gibba (contrib.%=4.77), ampelisca spinipes (contrib.%=3.80) and nucula sulcata (contrib.%=3.10), which mainly characterised inshore stations, while polychaetes like tharyx spp. (contrib.%=1.71) and sternaspis scutata (contrib.%=1.42) proliferated offshore (tab. 7). the results of the two-way crossed anosim test performed on abiotic parameters showed that the environmental assets of the investigated stations did not vary from september to march (r=0.332), whereas in both sampling periods a significant variation was detected between the cluster that included stations at <20 m depth and that with stations at >20 m depth (tab. 5). in the pca two principal components (factors) were identified (eigenvalue >1) which together explained 73.72% of the total variance, whereas the first and second factors explained 56.07% and 17.65% of the total variance, respectively (fig. 4a). sand percentage, depth, din and toc mostly contributed to the first factor, whereas chl a was the most important contributor to the second factor. tab. 4. diversity indexes for benthic diatoms and macrofauna in september 2010 and march 2011. benthic diatoms macrofauna d h’(loge) λ d h’(loge) λ september 2010 b02 1.787 2.624 0.093 5.091 2.337 0.169 b03 1.285 2.229 0.136 3.382 2.415 0.125 b04 0.734 1.074 0.533 2.304 2.284 0.112 c01 0.951 2.026 0.156 6.082 1.994 0.278 c02 1.538 2.477 0.105 7.144 2.050 0.299 c03 1.083 2.057 0.159 3.338 2.335 0.149 c04 0.663 0.580 0.770 2.700 2.292 0.141 d02 2.207 2.782 0.077 6.561 2.520 0.199 d03 0.856 1.237 0.463 3.454 2.328 0.160 d04 0.528 1.054 0.510 2.190 1.636 0.360 march 2011 b02 1.271 2.070 0.193 5.984 2.829 0.113 b03 0.917 1.699 0.257 2.818 2.332 0.129 b04 0.702 0.929 0.595 2.495 2.133 0.181 c01 0.692 0.310 0.891 5.788 2.866 0.105 c02 1.219 1.859 0.253 5.297 2.583 0.139 c03 1.083 1.289 0.498 2.651 2.266 0.145 c04 1.078 1.486 0.387 2.492 2.055 0.186 d02 1.139 1.598 0.334 5.921 2.626 0.137 d03 0.923 1.766 0.245 3.914 2.630 0.102 d04 1.289 1.141 0.574 4.488 2.760 0.110 d, richness; h’, diversity; λ, dominance. no nco mm er cia l u se on ly 69 a. franzo et al. since the anosim tests revealed that the sampling period did not explain significant proportions of community variations, all of the samplings stations were plotted on the pca factor-plane 1x2, which grouped the stations according to their environmental conditions (fig. 4b). two distinct groups were recognized with inshore (<20 m) and offshore (>20 m) stations in the right and left half of the plot, respectively. discussion environmental characteristics of the sediments in the study area during both study periods, the main chemical parameters and the sediment grain-size varied remarkably within the 10-point-station grid, suggesting the presence of two different sedimentary matrices, which most likely led to the development of different benthic communities. mud prevailed at all stations, but the percentage of sand inshore was higher than the one offshore. as outlined by frontalini et al. (2011), this result suggests that the presence of more dynamic hydrological conditions at the inshore stations likely favoured the accumulation of poorly sorted and deposited fine-grained material offshore. as expected, the different grain-size had also an influence on the toc sedimentary content: the negative correlation between sand percentage and toc content in september (n=10; r=0.68, p<0.05) and march (n=9; r=-0.97, p<0.001), suggests that the muddier sediment offshore exerted a greater retention of sediment-bound organic matter, which in turn could have consequences on the food availability for benthic faunas (semprucci et al. 2010, 2013). in september, higher concentrations of din, negatively correlated with water depth (n=9; r=-0.82, p<0.01) were observed at the shallower stations than offshore. although correlations do not allow to infer about cause-effect relationships, the presence of a decreasing pattern of din pore water concentrations with increasing water depth in september, early observed by frontalini et al. (2011) for nutrient concentrations in the water column, suggests that eventual discharge activities or runoff occurring along the coasts in late summer (i.e., september) are rapidly diluted offshore. in march, the lack of such a coast-to-offshore pattern in din pore water concentrations could be probably due to the presence of more dynamic hydrological conditions in late winter, possibly caused by bad weather conditions, which could have favoured the spreading and dilution of the terrestrial outfalls even inshore. benthic communities this study was specifically designed in order to include not only locations along the strictly coastal belt but tab. 5. outputs of the two-way crossed anosim test performed using the sampling period (september vs march) and water depths (<20 m vs >20 m) as factors. statistically significant results according to clarke and gorely (2006) are underlined. two-way crossed anosim r r factor period depth benthic diatoms 0.334 0.835 meiofauna 0.153 0.46 macrofauna 0.611 0.976 abiotic parameters 0.332 0.673 tab. 6. species contributions to average dissimilarity between benthic diatom communities at the inshore (depth <20 m) and offshore stations (depth >20 m). cut-off percentage 50%. depth<20 m vs depth >20 m average dissimilarity=71.47 depth <20 m depth >20 m species av. abund av. abund av. diss diss/sd contrib% cum.% paralia sulcata 450 14150 5.04 1.69 7.05 7.05 skeletonema costatum 16725 675 4.05 0.99 5.67 12.72 rhopalodia gibba 2250 0 3.65 1.6 5.1 17.82 gyrosigma spencerii 1012.5 25 3.56 1.52 4.98 22.81 navicula spp. 1 1462.5 50 3.38 1.39 4.73 27.53 pinnularia spp. 562.5 425 3.11 1.21 4.35 31.89 diploneis bombus 150 825 3.04 1.09 4.25 36.14 gyrosigma acuminatum 262.5 600 2.89 1.21 4.04 40.19 nitzschia spp. 1875 150 2.89 1.21 4.04 44.23 nitzschia punctata 187.5 0 2.44 0.91 3.42 47.65 gyrosigma spp. 412.5 150 2.44 1.02 3.41 51.06 av. abund, average abundance; av. diss, average dissimilarity; contrib%, species contribution to the average dissimilarity between groups; cum.%, cumulative percentage. no nco mm er cia l u se on ly 70synoptic study of three benthic communities also offshore, i.e., even at >12 nmi from the coastline. this approach, although partially biased by pseudo-replication at the level of station, allowed us to provide some insights on the environmental status of a marine region generally neglected by monitoring plans. indeed, although offshore stations are included as putative targets for monitoring protocols of the msfd, there are still many knowledge gaps on these systems. in fact, according to the recently published viewpoint on the msfd complementary approach for the assessment of pressures, knowledge and data gaps in southern european seas (crise et al., 2015), most countries reported significant lack of data on biodiversity acquired offshore (i.e,. beyond 12 nmi seaward from the baseline). although a temporal variation was expected comparing the chosen sampling periods, the results of both the cluster analysis and the two-way crossed anosim suggest that the spatial variability among stations exceeded the temporal one, revealing the presence of different environmental contexts that, to some extent, were maintained over the seasons. mpb depend upon several factors, such as light penetration, sediment grain-size, coastal hydrodynamics, nutrient availability, grazing pressure, contamination and temperature. the synergistic effects of all these abiotic variables select the composition and abundance of the microphytobenthic community. since the 10-point-station grid covered a quite large area, with different physical features at each station, the resulting abundances were quite diverse among the studied stations. this was more pronounced in september, when the difference between values inshore and offshore reached up to one order of magnitude (9900±424 cells cm–3 at c01 vs 86400±16971 cells cm–3 at st. d02). the mpb abundance was particularly low at st. c01 most likely because of the high percentage of sand in this station (44.8%). in fact, coarse-grained sediments, which are typically characterised by high hydrodynamism with frequent sediment resuspension, do not allow the establishment of an epipelic mpb community. in more sublittoral areas, at water depths comprised between 18 and 20 m (i.e., st. b02, c02 and d02), muddier sediments associated with lower hydrodynamism enabled the presence of abundant microalgal assemblages, in particular in september. conversely, at the offshore stations (st. b03, b04, etc.), deeper depths and the consequent low light availability at the sea bottom probably limited the mpb proliferation. although generally benthic diatoms are characterised by relevant seasonal changes (summer vs winter assemblages, each characterised by typical genera and species, cibic et al., 2012), in this study significant temporal variability in the composition of microalgae was not highlighted (two-way crossed anosim analysis; r<0.5). interestingly, the discovery of obligate benthic diatoms living at depths of about 50 m substantially extends the known depth range of these primary producers, which, in turn, could hold significant implications for our knowledge and estimates of oceanic productivity and biogeochemical cycling. the depth and light limits of benthic microalgae have been, in fact, very poorly investigated to date: few measurements of benthic microalgal production below 20 m depth, very few records of benthic microalgal chl a (combined with microscopic observations of viable benthic diatom assemblages) at depths below 60 m, and only 2 reports from below 100 m worldwide were published before the year 2000 (cahoon, 1999). recently, the presence of live active diatoms in sediments at depths down to 191 fig. 4. pca ordination diagram based on the selected variables (a); toc, total organic carbon; tn, total nitrogen; din, dissolved inorganic nitrogen; phaeo, phaeopigments; mpb, microphytobenthos; meio, meiofauna; macro, macrofauna. scatter diagram plotting factors 1 and 2 of sampling stations (b). station acronyms as in fig. 1. s, september 2010; m, march 2011. no nco mm er cia l u se on ly 71 a. franzo et al. m, and at <0.03% surface incident radiation have been documented (mcgee et al., 2008). within the 10-pointstation grid, p. sulcata was the most abundant species both in september and in march, thus being apparently not influenced by temporal variations as already reported in a seven-year study carried out in the gulf of trieste (cibic et al., 2012). the absence of a relation between p. sulcata and light availability (mcquoid and nordberg, 2003), suggests that this species is able to survive in the darkness for a long time. this is in accordance with the present study since paralia was more abundant where the light availability was likely limited, i.e. at the offshore stations (depth >20 m). this hypothesis is partly supported by the positive correlation between paralia abundance and water depth in both periods (n=10, r=0.91, p<0.001 in september; n=10, r=0.84, p<0.01 in march). in this regard it is worth mentioning that previous studies have reported that, in sciaphilous conditions, the heterotrophic utilization of dissolved organic compounds by benthic diatoms is likely to be an important survival strategy when light levels are too low for photosynthesis (tuchman et al., 2006). in september, gyrosigma was quite abundant inshore especially at st. b02 and st. c02 (>4500 cells cm–3), where the highest din concentrations were observed. our results are in accordance with hunter (2007), who reported that this genus tends to proliferate in high organic content conditions and warm water temperature. moreover, our low gyrosigma abundances in march could be attributed to colder temperatures and the mixing of the water column as a consequence of more frequent stochastic events. with the exception of the gulf of trieste and the venice lagoon, only a few studies on natural microphytobenthic communities were carried out in the adriatic sea. totti (2003) assessed the influence of the po river outflow on the tab. 7. macrofauna species contributions to average dissimilarity between the two sampling periods and at the inshore (depth <20 m) and offshore stations (depth >20 m). cut-off percentage 50%. september vs march average dissimilarity=72.75 september march species av. abund av. abund av. diss diss/sd contrib% cum.% tharyxs pp. 36.00 1.67 2.22 1.11 3.05 3.05 corbula gibba 466.00 60.67 2.01 1.06 2.76 5.81 ampelisca spinipes 160.00 36.33 1.92 1.23 2.63 8.44 tellimya ferruginosa 53.67 50.33 1.88 1.17 2.58 11.02 turritella communis 33.00 39.33 1.79 1.31 2.46 13.48 caulleriella multibranchis 0.00 15.67 1.72 1.07 2.37 15.85 polynoinae indet. 0.00 11.33 1.67 1.03 2.30 18.15 cirratulidae indet. 0.33 7.00 1.47 0.98 2.02 20.17 hydrobia acuta 22.67 38.67 1.41 1.14 1.94 22.11 sigalionidae indet. 4.67 0.00 1.40 1.18 1.92 24.04 lumbrineris adriatica 23.67 12.33 1.35 1.31 1.86 25.90 labidoplax thomsoni 9.33 14.00 1.33 1.17 1.82 27.72 nucula sulcata 34.33 20.00 1.31 0.94 1.80 29.53 aphroditidae indet. 0.33 7.33 1.26 1.02 1.73 31.26 abra prismatica 9.67 0.33 1.18 1.31 1.62 32.88 paradoneis lyra 4.33 1.00 1.17 1.03 1.61 34.48 ampelisca spp. 13.00 5.67 1.13 1.05 1.55 36.03 sternaspis scutata 28.33 38.00 1.10 1.35 1.51 37.54 nephtys spp. 6.00 8.33 1.06 0.98 1.45 38.99 necallianassa acanthura 2.00 3.67 1.04 1.02 1.44 40.43 apseudes latreillei 82.67 3.33 1.04 0.86 1.43 41.86 cirratulus spp. 5.67 0.67 1.02 0.69 1.40 43.26 heteromastus filiformis 5.33 4.00 1.00 1.13 1.37 44.62 maldanidae indet. 7.33 7.67 0.99 0.99 1.36 45.99 anadara demiri 21.33 11.67 0.99 0.84 1.36 47.34 priapulida indet. 3.00 1.67 0.95 0.81 1.31 48.65 nephtys hombergi 5.00 1.00 0.95 0.74 1.31 49.96 glycera tridactyla 3.67 4.33 0.95 1.09 1.31 51.26 to be continued on next page no nco mm er cia l u se on ly 72synoptic study of three benthic communities microphytobenthic distribution and found that the plume, which determined marked coast-offshore gradients, influenced both the abundance and community structure. mpb were largely dominated by diatoms with both epipelic motile (pleurosigma spp., gyrosigma spp., navicula spp.) and non-motile forms (paralia sulcata). in accordance with our data, totti (2003) also reported that offshore sediments were characterised by the dominance of the centric diatom paralia sulcata. for both meiofauna and macrofauna, the methodological approach adopted should be taken into account when comparisons with other surveys are made. in the present study, the upper size limit to discriminate among these two communities was 1000 µm, whereas in other studies this limit is lowered to 500 µm (simonini et al., 2005a; simonini et al., 2005b; simonini et al., 2007; semprucci et al., 2010; frontalini et al., 2011). thus, the fraction of organisms with a body size varying between 500 µm and 1000 µm can fall within meiofauna or macrofauna according to the size limit used, consequently affecting the absolute abundances. although our meioand macrofauna densities were generally lower than those reported in previous works (danovaro et al., 2000; frontalini et al., 2011; semprucci et al., 2010), these communities followed the same spatial variability reported by the other authors, i.e. meiofauna abundance tended to increase offshore while macrofauna showed the opposite pattern. the meiofaunal community was dominated by nematodes and copepods, as observed in most coastal ecosystems, including the adriatic sea (balsamo et al., 2010). differently from benthic diatoms and macrofauna, the cluster analysis and the two-way crossed anosim outputs did not point out the presence of distinct assemblages according to the different depth. we infer that the level of tab. 7. continued from previous page. depth<20 m vs depth >20 m average dissimilarity=80.18 species depth <20 m depth >20 m av. abund av. abund av. diss diss/sd contrib% cum.% corbula gibba 658.33 0.00 3.83 4.54 4.77 4.77 ampelisca spinipes 243.33 1.39 3.05 2.70 3.80 8.58 nucula sulcata 67.92 0.00 2.49 2.45 3.10 11.68 turritella communis 86.67 2.50 2.16 1.66 2.69 14.37 maldanidae indet. 18.75 0.00 1.94 3.90 2.42 16.79 lumbrineris adriatica 42.50 1.67 1.84 2.05 2.29 19.08 tellimya ferruginosa 102.08 18.61 1.74 1.34 2.17 21.25 anadara demiri 41.25 0.00 1.73 1.45 2.16 23.41 apseudes latreillei 17.08 0.28 1.51 1.13 1.89 25.30 clymene palermitana 11.25 0.00 1.43 1.98 1.78 27.07 tharyx spp. 4.17 28.61 1.37 1.11 1.71 28.78 cylichna cylindracea 7.08 0.56 1.18 1.66 1.48 30.26 glycera tridactyla 8.33 1.11 1.14 1.58 1.42 31.68 sternaspis scutata 19.58 42.22 1.14 1.28 1.42 33.10 melinna palmata 12.08 0.00 1.08 0.89 1.35 34.45 pectinaria auricoma 31.67 0.28 1.08 1.11 1.34 35.79 heteromastus filiformis 8.75 1.94 1.07 1.55 1.33 37.13 golfingia spp. 6.25 0.28 1.07 1.52 1.33 38.46 leucothoe venetiarum 13.33 0.28 1.06 1.15 1.32 39.78 labidoplax thomsoni 9.58 13.06 1.04 1.40 1.30 41.07 ampelisca spp. 18.75 3.06 1.04 1.05 1.29 42.36 pseudoleiocapitella fauveli 12.08 0.00 1.00 1.13 1.25 43.62 tellina pulchella 7.92 0.00 1.00 1.16 1.24 44.86 abra prismatica 10.00 1.67 0.93 1.21 1.16 46.02 nephtys spp. 3.75 9.44 0.91 1.26 1.13 47.15 caulleriella multibranchis 1.67 11.94 0.90 0.96 1.12 48.26 polynoinae indet. 0.83 8.89 0.85 0.97 1.06 49.32 timoclea ovata 6.67 0.00 0.83 0.89 1.04 50.36 av. abund, average abundance; av. diss, average dissimilarity; contrib%, species contribution to the average dissimilarity between groups; cum.%, cumulative percentage. no nco mm er cia l u se on ly 73 a. franzo et al. taxonomic identification obtained for meiofauna may be not sufficient for detecting properly the changes induced by depth that likely were present within the 10-point-station grid. however, the different composition of the main groups observed among stations indicated the presence of a well developed assemblage, more diverse offshore compared to the poorly structured inshore community. higher numbers of taxa were observed, in fact, at the stations with a depth >20 m whereas the lowest value (4) was observed at st. c01 in both september and march. nematodes, recognised as organisms tolerant to a wide range of stressors (semprucci et al., 2015), exceeded 90% of the total inshore meiofaunal community. as indicated by din concentrations higher at stations nearby the coast, the benthic domain was likely subjected to nutrient loads from the land. going offshore, this influence decreased, favouring the development of less tolerant organisms such as copepods. these crustaceans, that showed higher abundances offshore, are recognized as sensitive for several environmental stressors, e.g., o2 depletion (danovaro et al., 2004). although a clear difference among sampling periods was highlighted for macrofauna (r=0.611), the different depth (20 m) seemed the main factor in determining the presence of distinct inshore and offshore communities as indicated by both the cluster analysis and the two-way crossed anosim (r=0.976). among macrofauna, molluscs constituted the most abundant group inshore, whereas annelids were more abundant offshore. the characteristics of the offshore sediments with high toc contents could have limited the abundance of suspension feeders such as bivalves, favouring at the same time infaunal benthos such as annelids (semprucci et al., 2010). bivalves were mainly represented by corbula gibba, while amphipods by ampelisca spinipes and both species were typical of the inshore stations. the bivalve c. gibba is considered as an indicator of environmental instability, low oxygen content and organic enrichment. this species tends, in fact, to form dense populations where the macrofaunal community is characterised by low diversity as in constantly or occasionally unbalanced environments (hrsbrenko, 2006). in contrast, the fossorial crustacean ampelisca acts as a bioturbator, which builds and irrigates its burrows and, in conditions of o2 depletion, creates a small oxidized cylinder within the sediment enhancing sea bottom oxygenation. therefore ampelisca is considered a structural taxon since, modifying the substratum, makes it more suitable for supporting other species. for this reason, ampelisca has been credited as a good bioindicator of environmental recovery. in recent years, a sensible reduction of c. gibba average number has been reported along the emilia romagna coastline, compared to the period of 19802000 (n’siala et al., 2008). furthermore, the increased abundances of ampelisca suggest a gradual amelioration of the environmental status of the italian coasts. the pca confirmed the presence of two distinct environmental contexts within which different benthic communities develop. the stations with depths <20 m have a higher percentage of sand and seem to be influenced by nutrient loads of terrestrial origin as confirmed by higher din and chl a values (in september). the latter, in fact, can generally provide a good indication of labile organic inputs to sediments (boon and duineveld, 1998). these environmental conditions likely influence the three benthic communities analysed, resulting in assemblages dominated by tolerant and opportunistic taxa (gyrosigma and c. gibba) and groups of organisms (nematodes). the stations with depths >20 m have muddier sediments that likely restrain more toc than inshore ones. these environmental features could enhance deposit feeders, such as macrofaunal annelids, for whom the accumulation of toc could represent a noteworthy food source. conclusions in the study area, the three benthic communities showed a clear depth-related distribution. the isobath of 20 m seems to influence the selection of distinct inshore vs offshore autotrophic and heterotrophic benthic communities. besides this purely physical feature, also the chemistry of sediments likely affects the studied communities. in fact, the gradual increase of the distance from the coastline likely determines a dilution effect of the inorganic loads from the land as indicated by the decreasing concentrations of din. however, with the exception of the dissolved inorganic forms of nitrogen, that could be considered a proxy of terrestrial organic inputs, our data did not highlight the presence of a clear anthropogenic impact on the investigated benthic communities. the co-occurrence of sensitive taxa such as meiofaunal copepods and ampelisca with opportunistic ones (as c. gibba) suggests that the benthic domain is only moderately impacted. it was beyond the aim of this study to estimate the benthic diatom community production. however, the presence of a viable benthic diatom community at depths of about 50 m extends the known depth range of these primary producers and holds significant implications for oceanic primary production and biogeochemical cycling. interestingly, sciaphilous conditions even at depths ~50 m enhanced the presence of diatoms able to switch from the autotrophic to heterotrophic metabolism as paralia sulcata. acknowledgements the authors would like to thank c. comici for tn and toc data, c. falconi and m. kralj for nutrient analysis, m. lipizer for pigment analysis, m. bussi for granulometric determination and d. borme for his help during sampling. no nco mm er cia l u se on ly 74synoptic study of three benthic communities the authors warmly thank the crew of the r/v-daphne and the environmental protection agency of emilia-romagna region for the logistical support. a special thanks goes to r. auriemma for his help with the statistical analysis. references albertelli g, covazzi-harriague a, danovaro r, fabiano m, fraschetti s, pusceddu a, 1999. differential responses of bacteria, meiofauna and macrofauna in a shelf area (ligurian sea, nw mediterranean): role of food availability. j. sea res. 42:11-26. balsamo m, albertelli g, ceccherelli vu, coccioni r, colangelo ma, curini-galletti m, danovaro r, d’addabbo r, de leonardis c, fabiano m, frontalini f, gallo m, gambi c, guidi l, moreno m, pusceddu a, sandulli r, semprucci f, todaro ma, tongiorgi p, 2010. meiofauna of the adriatic sea: present knowledge and future perspectives. chem. ecol. 26:45-63. barranguet c, 1997. the role of microphytobenthic primary production in a mediterranean mussel culture area. estuar. coast. shelf s. 44:753-765. bonaglia s, nascimento fja, bartoli m, klawonn i, brüchert v, 2014. meiofauna increases bacterial denitrification in marine sediments. nat. commun. 5:5133. boon ar, duineveld gca, 1998. chlorophyll a as a marker for bioturbation and carbon flux in southern and central north sea sediments. mar. ecol.-prog. ser. 162:33-43. cahoon lb, 1999. the role of benthic microalgae in neritic ecosystems. oceanogr. mar. biol. annu. rev. 37:47-86. castelli a, lardicci c, tagliapietra d, 2004. soft bottom macrobenthos. in: m.c. gambi, and m. dappiano (eds.), mediterranean marine benthos: a manual of methods for its sampling and study. biol. mar. medit. 11(suppl. 1):99-131. cibic t, blasutto o, falconi c, fonda umani s, 2007. microphytobenthic biomass, species composition and nutrient availability in sublittoral sediments of the gulf of trieste (northern adriatic sea). estuar. coast. shelf s. 75:50-62. cibic t, comici c, bussani a, del negro p, 2012. benthic diatom response to changing environmental conditions. estuar. coast. shelf s. 115:158-169. cibic t, facca c, 2010. microphytobenthos. in: g. relini (ed.), checklist della flora e della fauna dei mari italiani. parte ii. biol. mar. medit 17(suppl.1):754-800. clarke a, gorely rn, 2006. primer v6: user maual/tutorial. primer-e ltd, plymouth, uk. clarke kr, warwick rm, 2001. change in marine communities: an approach to statistical analysis and interpretation. primer-e ldt, plymouth, uk. colantoni p, fanzutti gp, marocco r, 1985. geologia della piattaforma adriatica. progetto finalizzato oceanografia e fondi marini, sottoprogetto risorse minerarie, rapporto tecnico finale, pp. 59-85. cozzi s, giani m, 2011. river water and nutrient discharges in the northern adriatic sea: current importance and long term changes. cont. shelf res. 31:1881-1893. creel l, 2003. ripple effects: population and coastal regions. population reference bureau, washington, dc, usa. crema r, castelli a, prevedelli d, 1991. long term eutrophication effects on macrofaunal communities in northern adriatic sea. mar. pollut. bull. 22:503-508. crise a, kaberi h, ruiz j, zatsepin a, arashkevich e, giani m, karageorgis ap, prieto l, pantazi m, gonzalez-fernandez d, ribera d’alcalà m, tornero v, vassilopoulou v, durrieu de madron x, guieu c, puig p, zenetos a, andral b, angel d, altukhov d, ayata sd, aktan y, balcıoğlu e, benedetti f, bouchoucha m, buia mc, cadiou jf, canals m, chakroun m, christou e, christidis mg, civitarese g, coatu v, corsinifoka m, cozzi s, deidun a, dell’aquila a, dogrammatzi a, dumitrache c, edelist d, ettahiri o, fonda-umani s, gana s, galgani f, gasparini s, giannakourou a, gomoiu mt, gubanova a, gücü ac, gürses ö, hanke g, hatzianestis i, herut b, hone r, huertas e, irisson jo, i̇şinibilir m, jimenez ja, kalogirou s, kapiris k, karamfilov v, kavadas s, keskin ç, kideyş ae, kocak m, kondylatos g, kontogiannis c, kosyan r, koubbi p, kušpilić g, la ferla r, langone l, laroche s, lazar l, lefkaditou e, lemeshko ie, machias a, malej a, mazzocchi mg, medinets v, mihalopoulos n, miserocchi s, moncheva s, mukhanov v, oaie g, oros a, öztürk aa, öztürk b, panayotova m, prospathopoulos a, radu g, raykov v, reglero p, reygondeau g, rougeron n, salihoglu b, sanchez-vidal a, sannino g, santinelli c, secrieru d, shapiro g, simboura n, shiganova t, sprovieri m, stefanova k, streftaris n, tirelli v, tom m, topaloğlu b, topçu ne, tsagarakis k, tsangaris c, tserpes g, tuğrul s, uysal z, vasile d, violaki k, xu j, yüksek a, papathanassiou e, 2015. a msfd complementary appraoch for the assessment of pressures, knowledge and data gaps in southern european seas: the perseus experience. mar. pollut. bull. 95:28-39. danovaro r, fraschetti s, belgrano a, curini-galletti m, vincx m, alberelli g, fabiano m, 1995. the potential impact of meiofauna on the recruitment of macrobenthos in a subtidal coastal benthic community of the ligurian sea: a field result, p. 115-122. in: a. eleftheriou, a.d. ansell and c.j. smith (eds.), biology and ecology of shallow coastal waters. olsen & olsen, fredenborg. danovaro r, gambi mc, manini e, fabiano m, 2000. meiofauna response to a dynamic river plume front. mar. biol. 137:359-370. danovaro r, gambi c, mirto s, sandulli r, ceccherelli vu, 2004. meiofauna. in: m.c. gambi and m. dappiano (eds.), mediterranean marine benthos: a manual of methods for its sampling and study. biol. mar. medit. vol 11(suppl 1):55-97. dexing j, junmin l, zhaodi c, shicheng l, 1985. marine benthic diatoms in china. china ocean press, beijing: 313 pp. european commission, 2008. directive 2008/56/ec of the european parliament and of the council establishing a framework for community action in the field of marine environmental policy (marine strategy framework directive). official journal, l164, pp. 19-40. fain amv, ogston as, sternberg rw, 2007. sediment transport event analysis on the western adriatic continental shelf. cont. shelf res. 27:431-451. frontalini f, semprucci f, coccioni r, balsamo m, bittoni p, covazzi-harriague a, 2011. on the quantitative distribution and community structure of meio and macrofaunal communities in the coastal area of the central adriatic sea (italy). environ. monit. assess. 180:325-344. germain h, 1981. flore des diatomées eaux douces et no nco mm er cia l u se on ly 75 a. franzo et al. saumâtres.[book in french]. société nouvelle des éditions boubée, paris: 444 pp. grasshoff k, ehrhardt m, kremling k, almgren t, 1983. methods of seawater analysis. verlag chemie, weinheim: 419 pp. hendey ni, 1976. an introductory account of the smaller algae of british coastal waters. otto koeltz science publishers, koenigstein: 317 pp. higgins rp, thiel h, 1988. introduction to the study of meiofauna. smithsonian institution press, washington dc: 488 pp. hrs-brenko m, 2006. the basket shell corbula gibba as a species resistant to environmental disturbaces. acta adriat. 47:49-64. hunter j, 2007. diatoms as environmental indicators: a case study in the bioluminescent bays of vieques, puerto rico. proceedings of the 20th annual keck symp. wooster, oh, usa. available from: http://keckgeology.org/files/pdf/ symvol/20th/puertorico/hunter.pdf larson f, sundbäck k, 2008. role of microphytobenthos in recovery of functions in a shallow-water sediment system after hypoxic events. mar. ecol.-prog. ser. 357:1-16. lorenzen c, jeffrey j, 1980. determination of chlorophyll in sea water. unesco technical paper in marine science 35. avallabile from: http://unesdoc.unesco.org/images/0004/ 000434/043485eb.pdf manoukian s, spagnolo a, scarcella g, punzo e, angelici r, fabi g, 2010. effects of two offshore gas platforms on softbottom benthic communities (northwestern adriatic sea, italy). mar. environ. res. 70:402-410. margalef r, 1986. ecologia. omega, barcellona: 951 pp. macintyre hl, richard jg, miller dg, 1996. microphytobenthos: the ecological role of the “secret garden” of unvegetated, shallow-water marine habitats. i. distribution, abundance and primary production. estuaries19:186-201. massamba n’siala g, grandi v, iotti m, montanari g, prevedelli d, simonini r, 2008. responses of a northern adriatic ampelisca-corbula community to seasonality and short-term changes in the po river. mar. environ. res. 66:466-476. mcgee d, laws ra, cahoon lb, 2008. live benthic diatoms from the upper continental slope: extending the limits of marine primary production. mar. ecol.-prog. ser. 356:103-112. mcquoid mr, nordberg k, 2003. the diatom paralia sulcata as an environmental indicator species in coastal sediments. estuar. coast. shelf s. 56:339-354. morri c, bellan-santini d, giaccone g, bianchi cn, 2004. principles of bionomy: definition of assemblages and use of taxonomic descriptors (macrobenthos. in: m.c. gambi and m. dappiano (eds), mediterranean marine benthos: a manual of methods for its sampling and study. biol.a mar. medit. 11(suppl.1):573-600. mozetič p, solidoro c, cossarini g, socal g, precali r, francé j, bianchi f, de vittor c, smodlaka n, fonda umani s, 2010.recent trends towards oligotrophication of the northern adriatic: evidence from chlorophyll a time series. estuar. coasts 33:362-375. nieuwenhuize j, maas yem, middelburg jj, 1994. rapid analysis of organic carbon and nitrogen in particulate materials. mar. chem. 45:217-224. pella e, colombo b, 1973. study of carbon, hydrogen and nitrogen determination by combustion-gas chromatography. mikrochim. acta 5:697-719. poulain p-m, kourafalou vh, cushman-roisin b, 2001. northern adriatic sea, p. 143-165. in: b. cushman-roisin, m. gačić, p-m. poulain and a. artegiani (eds)., physical oceanography of the adriatic sea. kluwer academic publisher. ravaioli m, alvisi f, menegazzo-vitturi l, 2003. dolomite as a tracer for sediment transport and deposition on the northwestern adriatic continental shelf (adriatic sea, italy). cont. shelf res. 23:1359-1377. ricard m, 1987. atlas du phytoplancton marin. 2. diatomophycées. [book in french]. centre national de la recherché scientifique, paris: 297 pp. round fe, crawford rm, mann dg, 1992. the diatoms. cambridge university press, avon: 747 pp. schlitzer r, 2015. ocean data view. available from: http://odv. awi.de semprucci f, boi p, manti a, covazzi-harriague a, rocchi m, colantoni p, papa s, balsamo m, 2010. benthic communities along a littoral of the central adriatic sea (italy). helgol. mar. res. 64:101-115. semprucci f, frontalini f, covazzi-harriague a, coccioni r, balsamo m, 2013. meioand macrofauna in the marine area of the monte st. bartolo natural park (central adriatic sea, italy). sci. mar. 77:189-199. semprucci f, losi v, moreno m, 2015. a review on italian research on free-living marine nematodes and the future perspectives on their use as ecological indicators (ecoinds). medit. mar. sci. 16:352-365. shannon ce, weaver w, 1949. the mathematical theory of communication. illinois press, urbana: 117 pp. sharp jh, 1974. improved analysis for “particulate” organic carbon and nitrogen from seawater. limnol. oceanogr. 19:984-989. shepard fp, 1954. nomenclature based on sand silt clay ratios. j. sediment. petrol. 24:151-158. simonini r, ansaloni i, bovini p, grandi v, graziosi f, iotti m, massamba n’siala g, mauri m, montanari g, preti m, de nigris n, prevedelli d, 2007. recolonization and recovery dynamics of the macrozoobenthos after sand extraction in relict sand bottoms of the northern adriatic sea. mar. environ. res. 64:574-589. simonini r, ansaloni i, bonvicini pagliai am, cavallini f, iotti m, mauri m, montanari g, preti m, rinaldi a, prevedelli d, 2005a. the effects of sand extraction on the macrobenthos of a relict sands area (northern adriatic sea): results 12 months post-extraction. mar. pollut. bull. 50:68-777. simonini r, ansaloni i, cavallini f, graziosi f, iotti m, massamba n’siala g, mauri m, montanari g, preti m, prevedelli d, 2005b. effects of long-term dumping of harbor-dredged material on macrozoobenthos at four disposal sites along the emilia-romagna coast (northern adriatic sea, italy). mar. pollut. bull. 50:1595-1605. simpson eh, 1949. measurement of diversity. nature 163:688. sundbäck k, linares f, larson f, wulff a, 2004. benthic nitrogen fluxes along a depth gradient in a microtidal fjord: the role of denitrification and microphytobenthos. limnol. oceanogr. 49:1095-1107. tomadin l, 2000. sedimentary fluxes and different dispersion mechanisms of the clay sediments in the adriatic basin. rend. fis. accad. lincei 11:161-174. tomas cr, 1997. identifying marine phytoplankton. academic press, san diego: 858 pp. totti c, 2003. influence of the plume of the river po on the disno nco mm er cia l u se on ly 76synoptic study of three benthic communities tribution of subtidal microphytobenthos in the northern adriatic sea. bot. mar. 46:161-178. traykovski p, wiberg pl, geyer wr, 2007. observations and modelling wave-supported sediment gravity flows on the po prodelta and comparison to prior observations from the eel shelf. continent. shelf res. 27:357-399. tuchmann nc, schollett ma, rier st, geddes p, 2006. differential heterotrophic utilization of organic compounds by diatoms and bacteria under light and dark conditions. hydrobiologia 561:167-177. utermöhl h, 1958. zur vervollkommnung der quantitativen phytoplankton-methodik.[article in german]. mitt. int. ver. theor. angew. limnol. 9:1-38. van heurck h, 1899. traité des diatomées. édité aux frais de l’auteur, anvers: 572 pp. witkowski a, lange-bertalot a, metzeltin d, 2000. iconographia diatomologica. 7. annotated diatom micrographs. arg gantner verlag kg: 925 pp. zonneveld kaf, versteegh gjm, kasten s, eglinton ti, emeis k-c, huguet c, koch bp, de lange gj, de leeuw jw, middelburg jj, mollenhauer g, prahl fg, rethemeyer j, wakeham sg, 2010. selective preservation of organic matter in marine sediments; processes and impact on sedimentary record. biogeosciences 7:483-511. no nco mm er cia l u se on ly layout 1 plastics in the ocean plastic is today recognized as the most abundant form of man-made debris in the sea (barnes et al., 2009), with the mass of land-based plastic waste entering the ocean recently estimated to be in the range 4.8 to 12.7 million metric tons per year (jambeck et al., 2015). early reports on the occurrence of plastics in the marine environment can be traced back to the 70s: carpenter and smith (1972) reported on the presence of an average 3500 plastic particles per square kilometer in the sargasso sea surface, collected through a neuston net in an area located far from obvious pollution sources from land. the authors speculated that the source of the particles could have been the dumping of waste from cities or by the cargo and passenger ships, given that some of the sampled areas were within major shipping lanes from europe to america. later on, gregory and ryan (1997) reported that plastics accounted for a significant proportion (from 60 to 80%) of the total debris encountered in the seas of the southern hemisphere. since then, a number of papers have increasingly documented the presence and spread of plastic debris across the marine environment (galgani et al., 2000; moore et al., 2001; moore et al., 2002; willis et al., 2017; worm et al., 2017). however, the threat of plastics to the marine environment has been ignored for a long time, and it is only in more recent years that its serious consequences for the ocean, the wildlife and the human health have started to being recognized (derraik, 2002). plastic debris occurs along the coastlines (browne et al., 2010), at the sea surface (law et al., 2014) and on the sea floor (stefatos et al., 1999; galgani et al., 1996; claessens et al., 2011; cau et al., 2017), and even in remote areas such as the open ocean surface far from land (cozar et al., 2014). plastic reaches the ocean both in the form of large visible debris (“macro”) that is larger than 1-5 mm, or in the form of small particles or fragments called “micro-plastics” (having dimensions typically <15 mm) (browne et al., 2010), despite globally accepted definitions for these categories are yet to be established. in a recent review that compared the methodologies used in 68 studies for the quantification of microplastics in the marine environment, hidalgo-ruz et al. (2012) highlighted that most of them reported two main size ranges of microplastics, 500 μm-5 mm and 1-500 μm, or fractions thereof that are retained on filters, confirming that there is still not a universally adopted size range to define advances in oceanography and limnology, 2017; 8(2): 199-207 review doi: 10.4081/aiol.2017.7211 this work is licensed under a creative commons attribution-noncommercial 4.0 international license (cc by-nc 4.0). surfing and dining on the “plastisphere”: microbial life on plastic marine debris grazia marina quero,1 gian marco luna2* 1stazione zoologica anton dohrn, 80121 villa comunale, naples; 2institute of marine sciences (cnr-ismar), national research council, largo fiera della pesca 2, 60125 ancona, italy *corresponding author: gianmarco.luna@an.ismar.cnr.it abstract plastic marine debris represents a global threat for the marine environment, having serious consequences for the ocean, the wildlife and the human health. while the plastics distribution, fate, persistence and toxicity mechanisms for the marine fauna have been more studied in the last decade, small efforts have been devoted to identify and characterize marine microbes that colonize plastic and microplastic debris in the ocean, and their potential to degrade plastics. here we review the knowledge on the microbial biodiversity and degradation mechanisms of marine plastic debris, and present data, based on metagenomic analyses, on the distribution patterns of genes potentially involved in microbially-mediated plastic degradation in coastal locations across the global ocean. most studies on plastic-colonizing microbes have focused on seawater rather than sediment, with most studies underlining striking differences in composition between assemblages attached to plastic particles and those in the surrounding environment. the diversity of microbes attached to plastic is high, and the core epiplastic microbial assemblages include often hydrocarbon-degrading bacteria, as well as prokaryotic and eukaryotic phototrophs. several marine microbes have shown to be able to degrade or deteriorate plastic in the laboratory, or to grow on plastic as the only source of carbon, while indirect evidences suggest that microbially-mediated degradation of recalcitrant plastics also occur in the ocean, though at very low rates. metagenomic analyses show that plastic degradation-related genes are present in microbial assemblages in several coastal ocean sites, with relative abundance related to the magnitude of plastic pollution at each site. further research is required to study microbial plastic-degraders in the marine ecosystem, to decipher and exploit the potential of microbial consortia to degrade or mineralize plastic compounds, and to better understand the fate and residence times of plastic waste in the ocean. key words: plastics; microbes; metagenomics; biodegradation. received: december 2017. accepted: december 2017. no nco mm er cia l u se on ly g.m. quero and g.m. luna200 microplastics. thus, the term “microplastics” is used in the literature to include a surprisingly broad range of particles sizes from ~5 mm to few microns in diameter (andrady 2017). a more practical scheme to classify plastics debris in the aquatic environment has been proposed by the european msfd technical subgroup on marine litter (reported in van cauwenberghe et al., 2015), a scheme which also includes the category “mesoplastics” in the size range 5 mm-2.5 cm. plastic debris enters the marine environment in a wide range of sizes, in the micrometer to meter range (hidalgo-ruz et al., 2012), as well in a broad range of color, shape, chemical composition and specific gravity. as far as the small size plastic debris is concerned, microplastics are typically categorized into primary and secondary (andrady, 2017). the primary ones are purposefully created, and are directly released into the environment as small plastic particles. they are typically industrially manufactured, such as microbeads used in cosmetics, personal care, abrasive media and other industries, or as virgin plastics pellets for fabrication of products, and enter the environment via leakage during manufacture, transportation or use. a recent study estimated that between 15 and 31% of all of the microplastics in the oceans could originate from primary sources, and estimated the global release of primary microplastics into the ocean at 1.5 million tons per year (boucher and friot, 2017). the secondary microplastics, that are believed to be far more abundant in the ocean, originate from fragmentation of larger plastic debris items either during their use of products, or due to weathering degradation of their litter in the environment. examples of secondary microplastics include the textile fiber fragments released from synthetic fabrics during washing, degrading of agricultural mulch films, and the weathering breakdown of plastic litter in the marine environment (andrady, 2017). however, still different definitions for primary and secondary microplastics exist in the literature (boucher and friot, 2017). this holds especially true for those microplastics that originate from the abrasion of larger plastic objects during manufacturing, use or maintenance (such as the erosion of tyres when driving, that account for a significant source of plastics in the ocean) that are classified as primary by some authors (sundt et al., 2014; boucher and friot, 2017) or secondary by others (gesamp, 2015). sources of plastics and microplastics in the ocean are multiple, and the estimates of pollution sources and sinks are still uncertain. once plastic items reach the marine environment (fig. 1), positively buoyant ones will tend to accumulate at the sea surface, where they are transported by winds and surface water currents over long distances, whereas negatively buoyant items sink out of the water column to the sediments below (clark et al., 2016). it has to be noted that, among the classes of plastic that are commonly encountered in the marine environment, only a few have specific gravity lower than that of seawater (such as polyethylene, pe and polypropylene, pp) and are thus positively buoyant, while many of them (such as polyvinyl chloride, pvc) are denser and tend to submerge in the water column (andrady, 2011). because of fouling by micro(lobelle and cunliffe, 2011) and macro-organisms and the adherence of particles, positively buoyant plastics can, over a timescale of weeks to months, become negatively buoyant and sink to the seafloor down to the deep sea (woodall et al., 2014). plastic fragments are considered to be quite stable and highly durable (sivan 2011), potentially lasting hundreds to thousands of years in the aquatic environment (barnes et al., 2009). however, the exposure of plastic objects on the surface waters to uv radiation from sunlight results typically in their photodegradation by oxidation (gewert et al., 2015) (fig. 1). further degradation, fragmentation and erosion of the plastic debris are caused by physical forces (such as wave action, wind and abrasion with sand), biological breakdown and chemical weathering. during fragmentation and weathering of plastic debris in seawater, nano-scale particles are also generated (gigault et al., 2016), although their abundance has not been quantified yet in the global ocean (ter halle et al., 2017). this finding calls for further investigations on the fate, behavior and hazard posed by nano-plastics in the marine environment. in addition to the concerns raised by toxicity of the plastic polymers and their additives (often intentionally added during manufacturing or processing to improve performance; gewert et al., 2015) for marine life (through ingestion and other mechanisms), plastics and microplastics also adsorb and accumulate toxic chemicals, such as persistent organic pollutants (bakir et al., 2012), transporting them across the ocean and vehiculating to the marine organisms that ingest them (mato et al., 2001). recent findings showed that organisms belonging to three different phyla that live on the deep-sea floor do ingest and internalize plastic microfibers, demonstrating the even the deep sea and its inhabitants are being exposed to this type of anthropogenic waste (taylor et al., 2016). compared to mesoand macroplastic debris, microplastic is more prone to adsorb waterborne contaminants due to its higher surface area to volume ratio (gewert et al., 2015). recent studies have also paid attention to the concept of plastic providing a novel means of transport for microorganisms, including pathogenic ones, across the marine environment, consequently acting as a possible vector for the spread of these microbes that can facilitate the diffusion of infectious diseases (keswani et al., 2016). here, we review the current knowledge on microbes that are associated with plastic debris in the ocean. while plastic debris can be colonized by a wide variety of microbes, belonging to the three domains of life, including microbial eukaryotes (such as diatoms; oberno nco mm er cia l u se on ly microbial life on plastic debris 201 beckmann et al., 2016), we focus our review on the prokaryotic (bacteria and archaea) plastic colonizers. also, given that excellent reviews have already covered the type and biodiversity of microbes that colonize plastic and microplastic fragments in aquatic ecosystems (oberbeckmann et al., 2015; harrison et al., 2018), we focus here on summarizing the knowledge on the identity of those marine bacteria, and their genetic pathways, that colonize and potentially degrade (i.e., cause the plastic chemical breakdown) or deteriorate (i.e., cause loss of physical integrity; rummel et al., 2017) plastic debris. lastly, based on metagenomic analyses, we present data about the distribution patterns of some genes that are potentially involved in microbially-mediated plastic degradation in several coastal locations across the global ocean. fig. 1. schematic drawing that summarizes the fate of plastic debris in the marine environment, with a focus on the interactions between plastics and microbes. plastic debris reaches, mostly from land-based sources, the marine environment in the form of visible debris and of smaller pieces that are not visible to the naked eye (microplastics). once reached the ocean, floating plastic is readily colonized by a wide variety of microbes, including members from the three domains of life (archaea, bacteria and eukarya). marine viruses can likely be members of the epiplastic assemblages, but reports on this biotic component on plastic are still not available in the literature. the hydrophobic nature of plastic surfaces stimulates rapid formation of the microbial biofilm, which then drives succession of other microand macro-organisms (among which diatoms and invertebrates; reisser et al., 2014). this biofilm contributes significantly to determine the subsequent fate of the plastic items (down to the millimeter-size), by influencing their ballasting ability, by likely mediating their degradation and fragmentation to smaller particles (that is believed to occur through a combination of physical, chemical and biological processes) and by affecting the toxicity level. fragmentation produces smaller and smaller plastic fragments, down to the nanometers size. transport of plastic items to the seafloor occur through several ways, although the transport mechanisms of biofouled microplastics to the ocean interior still remain unclear. a recent model that explains the sinking vertical mechanisms of microplastic has been published by kooi et al. (2017). once reached the seafloor, the plastic debris will undergo colonization by the sediment microbes (harrison et al., 2014), degradation and/or permanent sequestration. however, sediments will not necessarily represent the ultimate sink for plastic debris, because of the removal and/or digestion by benthic animals living in the sediments (rummel et al., 2017). no nco mm er cia l u se on ly g.m. quero and g.m. luna202 microbial life on plastic debris once in the marine environment, microbes readily (i.e., within minutes to hours) colonize the aquatic plastic debris, and the members of these biofilm communities can include pathogenic, toxic, invasive or plastic degrading-species (oberbeckmann et al., 2016). to date, the largest fraction of studies on plastic-colonizing microbes have focused on seawater samples (dang and lovell, 2000; lee et al., 2008; carson et al., 2013; zettler et al., 2013; amaral-zettler et al., 2015; bryant et al., 2016), with most of them underlining the striking differences in the community composition of bacterial taxa on plastic particles with respect to the surrounding water column (zettler et al., 2013; bryant et al., 2016; oberbeckmann et al., 2016). moreover, some of those studies have also used shotgun metagenomic to perform microbial functional gene comparisons, and have showed that plastic-inhabiting microbes have a gene repertoire that is enriched for traits necessary for a surface-attached lifestyle (bryant et al., 2016). conversely to seawater, only little information is available for plastic-associated communities in sediments (harrison et al., 2014; de tender et al., 2015; de tender et al., 2017). in seawater, early studies on microbial life attached to plastic surfaces in the ocean reported that typical early colonizers are members of the gammaproteobacteria and alphaproteobacteria, with an increasing trend in bacteroidetes during time (dang and lovell, 2000; lee et al., 2008), the so-called “primary” and “secondary colonizers”, respectively (lobelle and cunliffe, 2011; oberbeckmann et al., 2015). these findings were confirmed by dang et al. (2008) who identified early biofilm communities on pvc samples as belonging to the order of rhodobacterales (alphaproteobacteria), in particular to the clade roseobacter. the first study that investigated microbial biofilms on plastic marine debris using modern techniques of massive dna sequencing was carried out by zettler and colleagues (2013). this study showed a huge diversity within plastic-associated microbial communities in the north atlantic subtropical gyre, and provided the first detailed picture of the microbial epiplastic assemblages (that the authors termed “plastisphere”) attached to fragments of pp and pe, which appeared to be highly represented by prokaryotic and eukaryotic phototrophs (the cyanobacteria phormidium and rivularia, as well as numerous diatoms such as navicula, nitzschia, sellaphora, stauroneis and chaetoceros) and typical free-living bacterial taxa (e.g., pelagibacter). they also highlighted a remarkable presence of other protists and showed, by combining sem and ngs, a colonization by stalked suctorian ciliates covered with bacteria. moreover, they also identified a “core” plastisphere community that included hydrocarbon-degrading bacteria (phormidium sp., pseudoalteromonas sp. on pp and pe fragments, and hyphomonadaceae on pmd and pp), pointing out a possible role of plastic-inhabiting microbes in the degradation of those polymers. oberbeckmann et al. (2016) observed the changes in plastic-associated microbial communities after deploying pet drinking bottles in different stations and seasons in the north sea. although abundant pet-colonizing taxa belonged to the bacterial phylum bacteroidetes (e.g., flavobacteriaceae, cryomorphaceae, saprospiraceae) and to the diatoms coscinodiscophytina and bacillariophytina, pet-colonizing prokaryotic and eukaryotic communities varied greatly with space and time. these results, and the comparison of pet communities with glass-colonizing communities in the same experiment, suggested that microbial assembly on plastics is both driven by conventional marine biofilm processes (thus, not necessarily by a selection of plastic-specific colonizers), although a small proportion of taxa (members of the cryomorphaceae and alcanivoraceae) were significantly discriminant of pet, leading to speculate that these groups may directly interact with the pet substrate. in the recent study by bryant et al. (2016) in the north pacific subtropical gyre, plastic-attached communities were shown to be dominated by few metazoan taxa and by several photoautotrophic and heterotrophic protists and bacteria. bryozoa, cyanobacteria, alphaproteobacteria and bacteroidetes dominated on all the plastic items. as similarly observed for bacterial taxa by zettler et al. (2013), with the only major exception of vibrionaceae, and confirmed by bryant et al. (2016) after reannotation of their amplicon data, the authors identified the most abundant prokaryotic groups as cyanobacteria (belonging to genera phormidium, rivularia and leptolyngbya) and alphaproteobacteria (with rhodobacteraceae and hyphomonadaceae as the most represented families), followed by flavobacteriia, cytophagia, sphingobacteriia, gammaand deltaproteobacteria. in only one sample, the bacteroidetes genera tunicatimonas and tenacibaculum made up approximately 10% of one sample. only few, recent studies described the composition of microbial assemblages on plastic items retrieved from sediments. in a two-weeks experiment of exposure of microplastics in sediments of the north sea, harrison et al. (2014) found that the dominant colonisers on ldpe (low density polyethylene) microplastic belonged to the genera arcobacter and colwellia. seafloor plastic samples collected in the north sea by de tender et al. (2015) showed a wide variation in bacterial community composition, with a dominance of alphaand gammaproteobacteria and bacteroidetes. de tender et al. (2017) explored the taxonomic composition of bacterial (and fungal) communities on pe plastic sheets and dolly ropes during long-term exposure on the seafloor, at a harbor and at an offshore location in the north sea. bacterial communities on pe sheets in the harbor changed over time, with a gradual decrease in the relative abundance of primary colonizers (alphaand beta-proteobacteria) and an increase in secondary colno nco mm er cia l u se on ly microbial life on plastic debris 203 onizers (bacteroidetes), according to previous studies (dang and lovell, 2000; lee et al., 2008; lobelle and cunliffe, 2011; oberbeckmann et al., 2015). the core bacteriome included the genera arenicella and methylotenera, in addition to sulfurovum and maritimimonas in the early stage, robiginitomaculum in the middle stage, and sulfitobacter and psychroserpens in the late stage of exposure. a different picture was observed in the offshore waters, where the bacterial community was dominated by flavobacteria and gammaproteobacteria, and no temporal changes could be observed. in the offshore, the core microbiome included bacteroidetes (flavobacteriaceae) and proteobacteria (caulobacterales, hyphomonadaceae rhodobacteraceae and alcanivoracaceae). plastics degradation by marine microbes owing to the slow rates of degradation and breakdown, marine plastics can persist long in the environment; however, real data on the kinetics of mineralization of plastics in the marine environment are still virtually not existent (andrady, 2011). rates of conventional plastics (i.e., those plastics derived from petrochemicals) degradation by bacteria performed even in optimized laboratory conditions are extremely slow (reviewed in krueger et al., 2015), indicating that plastics is highly recalcitrant to biodegradation. recent studies, based on scanning electron observations of the surface of plastic particles, have pointed out the potential role of marine bacterial populations in contributing to plastic degradation, thus potentially intervening in the fragmentation dynamics (zettler et al., 2013). this hypothesis is based on the evidence that pits visualized in the plastic debris conformed to bacterial shapes, thus suggesting an active hydrolysis of the hydrocarbon polymer. this finding was later confirmed by reisser et al. (2014) in another study carried out in australia in a wide number of coastal and oceanic plastic samples. the more recent evidences for plastic degradation and assimilation by the bacterium ideonella sakaiensis, that is able to efficiently convert pet into its two environmentally benign monomers (terephthalic acid and ethylene glycol) (yoshida et al., 2016), are providing further impetus to perform research in this area, and are suggesting the ecological and biotechnological importance of exploiting novel polymer-degrading taxa that are likely to populate the terrestrial and aquatic ecosystems. the ability of microbes to degrade plastic is known since several decades, with the first studies reporting the ability of certain bacteria and fungi to degrade plastic polymers (harvey et al., 1949; booth et al., 1968). insights into the distribution, identity and potential of plastic-degrading microbes in the marine environment have been later reported over the following decades. most of those studies have been based on the isolation and in vitro testing of single strains isolated from the marine environment and the nearby areas. balasubramanian et al. (2010) recently isolated, by enrichment techniques from samples consisting of degraded polyethylene and soil in plastic waste dumped sites in the gulf of mannar (india), bacteria (identified as arthrobacter sp. and pseudomonas sp.) that efficiently degrade high-density polyethylene (hdpe). several comprehensive reviews have been focused on the microorganisms and their enzymes that are able to degrade petroleum-based plastic polymers (shah et al., 2008; sivan, 2011), or on the types and mechanisms of degradation, including biodegradation, and the factors that influence these processes (singh and sharma, 2008). krueger et al. (2015) have provided a comprehensive overview of the current knowledge on the enzymes implicated in the biodegradation of conventional plastics, as well as on the mechanisms of biodegradation of several of the most common plastic types, both hydrolysable and non-hydrolysable. they also state that, despite the increasing amount of studies performed in the last decades, information about biodegradation and the biochemical breakdown of synthetic, persistent conventional plastics is still scarce. plastic degradation by microbes is believed to occur through different pathways, depending also on the type of polymer. restrepò-flòrez et al. (2014) recently provided a review on the microbial degradation of pe, highlighting that biodegradation of this type of plastic is complex and not fully understood, and that its biodegradation in the environment maybe a cooperative process, mediated by different interacting types of microorganisms. while several studies have been performed in vitro following isolation of plastic degraders, studies of microbial degraders of plastic items performed directly in the marine environment are rarer. in a recent study, bryant et al. (2016) identified, in the north pacific gyre, a number of putative xenobiotic biodegradation genes that were significantly more abundant on the plastic-attached microbes rather than on the non-attached microbes that lived in the surrounding seawater. these genes included including homogentisate 1,2-dioxygenase, n-ethylmaleimide reductase, a cytochrome p450 and 2,4-dichlorophenol 6-monooxygenase. in particular, homogentisate 1,2-dioxygenase is an enzyme that has been implicated in the degradation of polycyclic aromatic hydrocarbons, as well as styrene, while 2,4dichlorophenol 6-monooxygenase is a hydroxylase involved in the degradation of chlorinated aromatic pollutants. the genes encoding the two subunits of protocatechuate 3,4-dioxygenase, an aromatic-ring-cleaving enzyme implicated in lignin degradation, were also observed in plastic metagenome samples. while the authors could not discern whether these genes serve for microorganisms residing on plastic debris to degrade plastics, or to co-metabolize the adsorbed pollutants, this study is the first that reported candidate genes that are likely involved in plastic degradation in the oceanic environment. no nco mm er cia l u se on ly g.m. quero and g.m. luna204 to verify the presence and prevalence of plastic degradation-related genes in marine microbial assemblages, we looked for the presence and prevalence of go (gene ontology) terms and kegg enzymes corresponding to these genes (more details in supplementary tab. 2) within a set composed by 13 selected metagenomes that are made available by the ocean sampling day initiative (kopf et al., 2015), a project that simultaneously analyzed microbial community composition and functional traits in more than one hundred sites around the global coastal ocean (fig. 2a). fig. 2. a survey of the presence and the prevalence, within marine metagenomes from coastal sites, of microbial genes that are potentially associated with plastic degradation. a) location of the 13 sampling sites in the coastal ocean considered in our study. b) and c) presence and distribution of the gene “homogentisate 1,2-dioxygenase” and “protocatechuate 3,4-dioxygenase” in the different sites, reported as number of observations of the corresponding go terms (b) and kegg enzyme (c). the metagenomes here analyzed are taken from the ocean sampling day (osd) database available here: https://www.ebi.ac.uk/metagenomics/projects/erp009703. “interpro matches” and “complete go annotation” files were downloaded from the following link: https://www.ebi.ac.uk/metagenomics/projects/erp009703. the osd site original name were here changed for facility of reading as follows: osd4, naples; osd5, crete; osd7, papeete; osd28, bengala gulf; osd37, miami; osd56, hawaii; osd69, marghera port; osd95, singapore; osd96, azores; osd106, iceland; osd111, rio de aveiro; osd123, shikmona; osd124, osaka. more details about the samples, including information about the total mg reads number and the total go annotated terms per each of the samples, are provided in supplementary tab. 1. the map and barplots were obtained with the ‘maps’ (https://cran.r-project.org/package=maps), ‘ggplot2’ (http://ggplot2.org), ‘ggrepel’ (http://github.com/slowkow/ ggrepel) and ‘ggplot2’, ‘reshape2’ (http://www.jstatsoft.org/v21/i12/paper) and ‘rcolorbrewer’ (https://cran.r-project.org/package= rcolorbrewer) packages, respectively, using the r programming language and environment (r core team, 2017). no nco mm er cia l u se on ly microbial life on plastic debris 205 the results are shown in fig. 2b (for gos) and 2c (for kegg). the results showed that neither kegg enzymes nor go terms related to 2,4-dichlorophenol 6-monooxygenase, n-ethylmaleimide reductase and cytochrome p450/nadph-cytochrome p450 reductase were observed throughout the dataset. on the other hand, kegg enzymes and go terms observations for homogentisate 1,2-dioxygenase-related counts ranged from 4 (azores) to 92 (hawaii) while, for the protocatechuate 3,4-dioxygenase, these were only observed as gos in the samples from naples (n=2), miami (n=4), hawaii (n=1) and singapore (n=1) and, as kegg enzymes, ranged from 0 (crete and osaka) to 13 (naples). homogentisate 1,2-dioxygenase-related go terms and enzymes were found in all the considered samples, with the highest number observed in sites such as hawaii, miami, naples and shikmona (fig. 2b and c). these sites are known in the literature to be particularly polluted by plastics (eriksen et al., 2014; van der hal et al., 2017). protocatechuate 3,4-dioxygenase related go terms were found in hawaii, miami, naples and singapore (fig. 2b). kegg enzymes codes for the same function were retrieved in all the samples except for crete and osaka, with the highest number of observations observed in hawaii, naples, miami, marghera and iceland (fig. 2c). although it is not possible to link directly these metagenomic data with plastic degradation processes occurring in the studied sites, these results suggest that degradation potential for plastic debris is ubiquitously present in marine microbial assemblages from almost all the coastal ocean, and that the relative abundance of microbial degrading genes could be related to the magnitude of plastic pollution within each site. further studies are required, combining omics techniques with in situ degradation rate measurements, to fully elucidate the potential of plastic breakdown of marine microbial assemblages. conclusions and future perspectives of research whether microbial plastic degradation or the use of plastic-associated chemicals are relevant processes in the oceans, and which environmental factors are able to influence these activities, are fundamental, yet still unanswered questions. little is known about the identity of microbes that colonize macroand microplastic debris in the ocean, and on their ability to degrade, transform or eventually assimilate plastics in the marine environment, including the time scales of degradation and the type and fate of the degradation products. a recent study has shown that tailored indigenous marine communities comprising polymer and hydrocarbon degrader species had the potential to degrade naturally weathered pe films, suggesting a more relevant role for microbial consortia, rather than single bacterial species, to degrade recalcitrant plastic (syranidou et al., 2017). the potential applications of the results gathered from researches on plastic-degrading microbially-mediated processes are not straightforward. in fact, the use of plastic-degrading microbes in the frame of large-scale applications to clean the ocean is obviously not a feasible strategy and, analogously, the use of plastic-degrading microbes is today not a reliable and effective alternative to conventional plastic recycling, that is based on plastic melting and regeneration. however, as the system of plastic recycling is still not fully exploited and optimized, the potential of microorganisms to degrade plastic has to be explored to learn more about degradation pathways, to ensure the safer disposal of plastic waste (via biodegradation), and to develop materials that decompose more readily than conventional plastic polymers. in addition, ecological research in this area is necessary to better understand the fate and the residence times of plastic waste in the ocean, and to predict the ecological consequences of plastic and microplastics as they are transported through the global ocean, and down to its interior (oberbeckmann et al., 2015). further microbiological research should be focused on studying plastic degraders in all marine habitats, including those habitats where plastic may accumulate (such as sediments), to exploit the potential of consortia to degrade and mineralize these compounds, and to combine in vitro and field studies with omics technologies to decipher the processes of microbial biodegradation. acknowledgments the work was made possible thanks to support granted to gml by the flagship programme ritmare (funded by the italian ministry of university and research), and to gmq by the szn and by the research prize “ricerca il futuro” funded by davines (www.ricercailfuturo.it). metagenomic data were provided by the ocean sampling day project (https://www.microb3.eu/osd.html). gml wish to thank his colleague giulio pellini for fruitful and stimulating discussions on this topic. references amaral-zettler la, zettler er, slikas b, boyd gd, melvin dw, morrall ce, et al. 2015. the biogeography of the plastisphere: implications for policy. front. ecol. environ. 13:541-546. andrady al, 2011. microplastics in the marine environment. mar. poll. bull. 62:1596-1605. andrady al, 2017. the plastic in microplastics: a review. mar. poll. bull. 119:12-22. bakir a, rowland sj, thompson rc, 2012. competitive sorption of persistent organic pollutants onto microplastics in the marine environment. mar. poll. bull. 64:2782-2789. no nco mm er cia l u se on ly g.m. quero and g.m. luna206 balasubramanian v, natarajan k, hemambika b, ramesh n, sumathi cs, kottaimuthu r, rajesh kannan v, 2010. highdensity polyethylene (hdpe)-degrading potential bacteria from marine ecosystem of gulf of mannar, india. lett. appl. microbiol. 51:205-211. barnes dka, galgani f, thompson rc, barlaz m, 2009. accumulation and fragmentation of plastic debris in global environments. philos. t. r. soc. lond. b biol. 364:1985-1998. booth gh, cooper aw, robb ja, 1968. bacterial degradation of plasticized pvc. j. appl. microbiol. 31:305-310. boucher j, friot d, 2017. primary microplastics in the oceans: a global evaluation of sources. iucn, gland: 43 pp. bryant ja, clemente tm, viviani da, fong aa, thomas ka, kemp p, karl dm, white ae, delong ef, 2016. diversity and activity of communities inhabiting plastic debris in the north pacific gyre. msystems 1:e00024-16. browne ma, galloway ts, thompson rc, 2010. spatial patterns of plastic debris along estuarine shorelines. environ. sci. technol. 44:3404-3409. carpenter ej, smith kl, 1972. plastics on the sargasso sea surface. science 175:1240-1241. carson hs, nerheim ms, carroll ka, eriksen m, 2013. the plastic-associated microorganisms of the north pacific gyre. mar. poll. bull. 75:126-132. cau a, alvito a, moccia d, canese s, pusceddu a, cannas r, angiolillo m, follesa mc, 2017. submarine canyons along the upper sardinian slope (central western mediterranean) as repositories for derelict fishing gears. mar. poll. bull. 123:357-364. claessens m, de meester s, van landuyt l, de clerck k, janssen cr, 2011. occurrence and distribution of microplastics in marine sediments along the belgian coast. mar. pollut. bull. 62:2199e2204. clark jr, cole m, lindeque pk, fileman e, blackford j, lewis c, lenton tm, galloway ts, 2016. marine microplastic debris: a targeted plan for understanding and quantifying interactions with marine life. front. ecol. environ. 14:317-324. dang h, li t, chen m, huang g, 2008. cross-ocean distribution of rhodobacterales bacteria as primary surface colonizers in temperate coastal marine waters. appl. environ. microbiol. 74:52-60. dang hy, lovell cr, 2000. bacterial primary colonization and early succession on surfaces in marine waters as determined by amplified rrna gene restriction analysis and sequence analysis of 16s rrna genes. appl. environ. microbiol. 66:467-475. de tender c, schlundt c, devriese li, mincer tj, zettler er, amaral-zettler la, 2017. a review of microscopy and comparative molecular-based methods to characterize “plastisphere” communities. anal. methods 9:2132-2143. de tender ca, devriese li, haegeman a, maes s, ruttink t, dawyndt p, 2015. bacterial community profiling of plastic litter in the belgian part of the north sea. environ. sci. technol. 49:9629-9638. derraik jg. 2002. the pollution of the marine environment by plastic debris: a review. mar. pollut. bull. 44:842-852. eriksen m, lebreton lc. carson hs, thiel m, moore cj, borerro jc, galgani f, ryan pg, reisser j, 2014. plastic pollution in the world's oceans: more than 5 trillion plastic pieces weighing over 250,000 tons afloat at sea. plos one 9:e111913. galgani f, souplet a, cadiou y, 1996. accumulation of debris on the deep sea floor off the french mediterranean coast. mar. ecol. progr. ser. 142:225-234. galgani f, leaute jp, moguedet p, souplet a, verin y, carpentier a, et al., 2000. litter on the sea floor along european coasts. mar. poll. bull. 40:516-527. gesamp, 2015. sources, fate and effects of microplastics in the marine environment: a global assessment. p.j. kershaw (ed.). imo/fao/unesco-ioc/unido/wmo/iaea/ un/unep/undp joint group of experts on the scientific aspects of marine environmental protection. rep. stud. gesamp no. 90: 96 pp. gewert b, plassmann mm, macleod m, 2015. pathways for degradation of plastic polymers floating in the marine environment. environ. sci-proc. imp. 17:1513-1521. gigault j, pedrono b, maxit b, ter halle a, 2016. marine plastic litter: the unanalyzed nano-fraction. environ. sci. nano 3:346-350. gregory mr, ryan pg, 1997. pelagic plastics and other seaborne persistent synthetic debris: a review of southern hemisphere perspectives, p. 49-66. in: j.m. coe and d.b. rogers (eds.), marine debris. springer, new york. harrison j p, schratzberger m, sapp m, osborn am, 2014. rapid bacterial colonization of low-density polyethylene microplastics in coastal sediment microcosms. bmc microbiol. 14:232. harrison jp, hoellein tj sapp m, tagg as, ju-nam y, ojeda jj, 2018. microplastic-associated biofilms: a comparison of freshwater and marine environments, p. 181-201. in: m. wagner, and s. lambert (eds.), freshwater microplastics. the handbook of environmental chemistry. springer, cham. harvey jv, meloro fa, 1949. studies of degradation of plastic films by fungi and bacteria. quartermaster laboratory report, microbiological series, washington dc. hidalgo-ruz v, gutow l, thompson rc, thiel m, 2012. microplastics in the marine environment: a review of the methods used for identification and quantification. environ. sci. technol. 46:3060-3075. keswani a, oliver dm, gutierrez t, quilliam rs, 2016. microbial hitchhikers on marine plastic debris: human exposure risks at bathing waters and beach environments. mar. environ. res. 118:10-19. kooi m, van nes eh, scheffer m, koelmans aa, 2017. ups and downs in the ocean: effects of biofouling on the vertical transport of microplastics. environ. sci. technol. 51: 963-7971. kopf a, bicak m, kottmann r, schnetzer j, kostadinov i, lehmann k, et al., 2015. the ocean sampling day consortium. gigascience 4:1-5. krueger mc, harms h, schlosser d, 2015. prospects for microbiological solutions to environmental pollution with plastics. appl. microbiol. biotechnol. 99:8857-8874. jambeck jr, geyer r, wilcox c, siegler tr, perryman m, andrady a, narayan r, law kl. 2015. plastic waste inputs from land into the ocean. science 347:768-771. law kl, morét-ferguson se, goodwin ds, zettler er, deforce e, kukulka t, proskurowski g. 2014. distribution of surface plastic debris in the eastern pacific ocean from an 11-year data set. environ. sci. technol. 48:4732-4738. lee j-w, nam j-h, kim y-h, lee k-h, lee d-h, 2008. bacterial communities in the initial stage of marine biofilm forno nco mm er cia l u se on ly microbial life on plastic debris 207 mation on artificial surfaces. j. microbiol. 46:174-182. lobelle d, cunliffe m, 2011. early microbial biofilm formation on marine plastic debris. mar. pollut. bull. 62:197-200. mato y, isobe t, takada h, kanehiro h, ohtake c, kaminuma t, 2001. plastic resin pellets as a transport medium for toxic chemicals in the marine environment. environ. sci. technol. 35:318-324. moore cj, moore sl, leecaster mk, weisberg sb. 2001. a comparison of plastic and plankton in the north pacific central gyre. mar. poll. bull. 42:1297-1300. moore cj, moore sl, weisberg sb, lattin gl, zellers af, 2002. a comparison of neustonic plastic and zooplankton abundance in southern california’s coastal waters. mar. poll. bull. 44:1035-1038. oberbeckmann s, loeder mg, gerdts g, osborn am. 2014. spatial and seasonal variation in diversity and structure of microbial biofilms on marine plastics in northern european waters. fems microbiol. ecol. 90:478-492. oberbeckmann s, löder mg, labrenz m, 2015. marine microplastic-associated biofilms-a review. environ. chem. 12:551-562. oberbeckmann s, osborn am, duhaime mb, 2016. microbes on a bottle: substrate, season and geography influence community composition of microbes colonizing marine plastic debris. plos one 11:e0159289. r core team, 2017. r: a language and environment for statistical computing. available from: https://www.r-project.org/ reisser j, shaw j, hallegraeff g, proietti m, barnes dka, et al., 2014. millimeter-sized marine plastics: a new pelagic habitat for microorganisms and invertebrates. plos one 9:e100289. restrepo-flórez j-m, bassi a, thompson mr, 2014. microbial degradation and deterioration of polyethylene-a review. intern. biodeterior. biodegradation 88:83-90. rummel cd, jahnke a, gorokhova e, kühnel d, schmittjansen m, 2017. the impacts of biofilm formation on the fate and potential effects of microplastic in the aquatic environment. environ. sci. technol. lett. 4:258-267. singh b, sharma n, 2008. mechanistic implications of plastic degradation. polymer degrad. stabil. 93:561-584. sivan a, 2011. new perspectives in plastic biodegradation. curr. opin. biotechnol. 22:422-426. shah aa, fariha h, abdul h, safia a, 2008. biological degradation of plastics: a comprehensive review. biotechnol. adv. 26:2467-2650. stefatos a, charalampakis m, papatheodorou g, ferentinos g, 1999. marine debris on the seafloor of the mediterranean sea: examples from two enclosed gulfs in western greece. mar. poll. bull. 36:389-393. sundt p, schulze p-e, syversen f, 2014. sources of microplastics-pollution to the marine environment. norvegian environ. agency, mepex 1032: pp. 86. syranidou e, karkanorachaki k, amorotti f, repouskou e, kroll k, kolvenbach b, et al., 2017. development of tailored indigenous marine consortia for the degradation of naturally weathered polyethylene films. plos one 12 e0183984. taylor ml, gwinnett c, robinson lf, woodall lc, 2016. plastic microfibre ingestion by deep-sea organisms. sci. rep. 6: 33997. ter halle a, jeanneau l, martignac m, jardé e, pedrono b, brach l, gigault j, 2017. nanoplastic in the north atlantic subtropical gyre. environ. sci. technol. 51:13689-13697. van cauwenberghe l, devriese l, galgani f, robbens j, janssen cr, 2015. microplastics in sediments: a review of techniques, occurrence and effects. mar. environ. res. 111:5-17. van der hal n, ariel a, angela dl, 2017. exceptionally high abundances of microplastics in the oligotrophic israeli mediterranean coastal waters. mar. pollut. bull. 116:151-155. willis k, hardesty bd, kriwoken l, wilcox c, 2017. differentiating littering, urban runoff and marine transport as sources of marine debris in coastal and estuarine environments. sci. rep. 7:44479. worm b, lotze hk, jubinville i, wilcox c, jambeck j, 2017. plastic as a persistent marine pollutant. ann. rev. environ. resour. 42:1-26. woodall lc, sanchez-vidal a, canals m, paterson gl, coppock r, sleight v, et al., 2014. the deep sea is a major sink for microplastic debris. royal soc. open sci. 1:140317. yoshida s, hiraga k, takehana t, taniguchi i, yamaji h, maeda y. toyohara k, miyamoto k, kimura y, oda k, 2016. a bacterium that degrades and assimilates poly (ethylene terephthalate). science 351:1196-1199. zettler er, mincer tj, amaral-zettler la, 2013. life in the “plastisphere”: microbial communities on plastic marine debris. environ. sci. technol. 47:7137-7146.no nco mm er cia l u se on ly layout 1 advances in oceanography and limnology, 2016; 7(1): 1-15 review doi: 10.4081/aiol.2016.5591 introduction the genus ostreopsis belongs to the family of ostreopsidaceae (gonyaulacales, dinophyceae, dinophyta). it was first described by schmidt (1901) after the finding of o. siamensis schmidt in plankton samples collected in the gulf of siam in 1900. since then, this species has rarely been encountered in the phytoplankton, due to its predominantly benthic habit. ostreopsis species were reported since a long time in tropical ciguatera endemic areas, associated with the benthic toxic dinoflagellate gambierdiscus toxicus (ballantine et al., 1985; carlson and tindall, 1985; bomber and aikman, 1989) and, therefore, improperly considered in association with ciguatera syndrome (tosteson, 1995). indeed, some ostreopsis species are toxic, but their toxins (mostly belonging to the palytoxin group) are not those implicated in ciguatera. in the last decade, ostreopsis blooms have become common in temperate areas as well, and regularly occur in the mediterranean sea during summer-autumn (vila et al., 2001; turki, 2005; aligizaki and nikolaidis, 2006; mangialajo et al., 2008; totti et al., 2010; illoul et al., 2012; ismael and halim, 2012; pfannkuchen et al., 2012), and in other temperate areas of the world (chang et al., 2000; rhodes et al., 2000; pearce et al., 2001; taniyama et al., 2003; shears and ross, 2009; selina et al., 2014). in these areas, ostreopsis is well-known since its blooms are often associated with noxious effects on health of both humans (gallitelli et al., 2005; kermarec et al., 2008; tichadou et al., 2010; del favero et al., 2012) and benthic marine organisms (pagliara and caroppo, 2012; gorbi et al., 2013; carella et al., 2015). additionally, ostreopsis often appeared in association with other toxic or potentially toxic benthic dinoflagellates such as prorocentrum spp., amphidinium spp. and coolia monotis in both several mediterranean (tognetto et al., 1995; vila et al., 2001; aligizaki and nikolaidis, 2006; monti et al., 2007; mabrouk et al., 2011) and world areas (okolodkov et al., 2007; parsons and preskitt, 2007; kim et al., 2011; selina and levchenko, 2011). the toxic benthic dinoflagellates of the genus ostreopsis in temperate areas: a review stefano accoroni,* cecilia totti dipartimento di scienze della vita e dell'ambiente, università politecnica delle marche, via brecce bianche, 60131 ancona, italy *corresponding author: s.accoroni@univpm.it abstract the genus ostreopsis includes species largely distributed from tropical to temperate marine areas worldwide. among the nine species of the genus, o. siamensis, o. mascarenensis, o. lenticularis and o. cf. ovata can produce toxins of the palytoxin group. in the last decade ostreopsis cf. ovata and o. cf. siamensis originated intense blooms in all the rocky mediterranean sea coastal areas, typically during summer-late summer. the correct identification of ostreopsis species in field samples is often problematic as ostreopsis species are morphologically plastic and hardly discriminable under light microscopy and, therefore, molecular analyses are required. ostreopsis blooms are often associated with noxious effects on health of both humans and benthic marine organisms mainly carried by aerosol and direct contact with seawater. environmental factors have been shown to affect toxin content of ostreopsis which generally produces more toxins per cell when growing under suboptimal conditions. o. cf. ovata is able to produce both temporary and resting cysts. in particular, the resting cysts are able to germinate in laboratory conditions for as long as 5 months after their formation at 25°c, but not at 21°c; the presence of a temperature threshold affecting cyst germination in the laboratory suggests that temperature represents a key factor for ostreopsis cf. ovata bloom onset in natural environments as well. several studies conducted to assess the role of abiotic factors (mainly hydrodynamics, water temperature and nutrients) on the bloom dynamics, revealed that the synergic effects of hydrodynamics, temperature and n:p ratios would lead the ostreopsis blooms in temperate areas. ostreopsis abundances showed a significant decrease with depth, likely related to light availability, although there are conflicting data about the relationship between light intensity and ostreopsis growth in experimental conditions. the relationship between ostreopsis blooms and salinity is not completely clear, complicated by the influence of high nutrient levels often associated to low salinity waters. finally, ostreopsis colonize a variety of substrata, although living substrata seems to allow lower concentration of epibionts than any other substrate, probably due to the production of some allelopathic compounds. key words: ostreopsis; palytoxin; harmful algae; benthic dinoflagellates; mediterranean sea. received: october 2015. accepted: february 2016. no nco mm er cia l u se on ly 2 s. accoroni and c. totti given their negative implications on both marine ecosystem functioning and human health and activities, ostreopsis blooms attracted the attention of researchers in the last decade. this paper is meant to be a narrative review of the existing information about taxonomy, geographical distribution, toxin production, life cycle and ecology of ostreopsis in temperate areas. taxonomy there is a considerable confusion regarding the descriptions of the ostreopsis species, since its first description carried out by schmidt (1901): in the original drawings of ostreopsis siamensis, both the epitheca and the hypotheca in anteroposterior view were presented but with marked different shapes (one appeared rounded and the other appeared elongated) insomuch hardly they can belong to the same cell. much later, fukuyo (1981) re-described o. siamensis and described two new species o. lenticularis and o. ovata. the rounded shape reported by schmidt for o. siamensis was one of the main differences between o. lenticularis and o. siamensis, so probably in the original description schmidt confused these two species. o. ovata was distinguished from the other two species by having a more ovoid shape and a smaller size. in the following years, several other species have been described by other authors: o. heptagona (norris et al., 1985), o. mascarenensis (quod, 1994), o. labens (faust and morton, 1995), o. marinus, o. belizeanus and o. caribbeanus (faust, 1999) (recently, the latter three have been renamed as o. marina, o. belizeana, o. caribbeana, hoppenrath et al., 2014). so, at present, nine species of ostreopsis have been described. in the mediterranean sea, only two species have been recorded until now, o. cf. ovata and o. cf. siamensis (vila et al., 2001; penna et al., 2005, 2010, 2012; battocchi et al., 2010; totti et al., 2010; mangialajo et al., 2011; perini et al., 2011; mabrouk et al., 2012). the correct identification of these ostreopsis species in field samples is often highly problematic. the most suitable taxonomical character used to discriminate o. cf. ovata and o. cf. siamensis, is the dorsoventral/anteroposterior diameter ratio (dv/ap) which is <2 and >4, respectively (penna et al., 2005; aligizaki and nikolaidis, 2006; selina and orlova, 2010). however, o. cf. ovata cells in the adriatic sea (mediterranean sea) have a dv/ap ratio slightly higher than 2 (i.e., 2.3-2.4) (monti et al., 2007; guerrini et al., 2010; accoroni et al., 2012b). since the molecular analyses clearly confirmed the presence of only o. ovata species in the n adriatic samples (perini et al., 2011), it can be hypothesized that the adriatic population has a different cell morphology, with cells more flattened than in other areas. these morphological problems and the lack of genetic data for the holotype specimens from which the original species descriptions of both o. siamensis and o. ovata were made, lead to the conclusion that o. siamensis and o. ovata examined nowadays should be referred to o. cf. siamensis and o. cf. ovata respectively, until more accurate morphological data and genetic sequences will be gathered to clearly define each species (penna et al., 2005, 2010). given these ambiguities in defining morphological characteristics, many researchers have been induced to revise the description of ostreopsis species by sequencing the its and 5.8s rdna regions, using these data in combination with morphometric ones. leaw et al. (2001) isolated several ostreopsis strains from malaysian coastal waters and showed that o. cf. ovata isolates were separated into two genetically distinct geographic groups, a malacca strait group and a south china sea group, while there were minor morphological differences among the strains. the 5.8s and its sequences of these malaysian strains differed from mediterranean strains ones (penna et al., 2005), suggesting a genetic variability in relation to the geographic distribution within the species. nowadays, the its-5.8s and lsu rdna allowed to distinguish various clades among the ostreopsis species (penna et al., 2014): the species complex ostreopsis cf. ovata includes the atlantic/mediterranean/pacific clade (i.e., isolates from japan sea, and mediterranean sea), the atlantic/indian/pacific clade (i.e., isolates from belize) and the pacific clade (i.e., isolates from vietnam); ostreopsis cf. siamensis forms an atlantic/mediterranean clade. the ostreopsis cf. lenticularis/o. cf. labens contains isolates from hawaii and pacific asia. distribution in temperate areas the first record of ostreopsis in the mediterranean sea dates back to 1972 in villefranche-sur-mer (france) by taylor (1979). later, o. cf. ovata was detected in 1994 along both the italian coasts of the tyrrhenian sea (tognetto et al., 1995) and the catalan coast in spain in 1997-1998 (vila et al., 2001). in the last decade, ostreopsis spp. blooms have been more intense, frequent, and widely distributed in many mediterranean areas, including spain, france, greece, italy, algeria, tunisia, turkey (turki, 2005; aligizaki and nikolaidis, 2006; ciminiello et al., 2006, 2008; riobó et al., 2006; turki et al., 2006; monti et al., 2007; riobó et al., 2008; guerrini et al., 2010; mabrouk et al., 2011, 2012; mangialajo et al., 2011; illoul et al., 2012). as reported above, genetic analyses indicate that two genotypes corresponding to the morphotypes o. cf. ovata and o. cf. siamensis are present in the mediterranean sea (penna et al., 2010, 2012). along the mediterranean rocky coasts, the genotype ovata is the most abundant and widely distributed (battocchi et al., 2010; perini et al., 2011). the genotype siamensis was detected along the catalan coast, no nco mm er cia l u se on ly ostreopsis in temperate areas in the eastern atlantic coast of morocco, portugal, northern spain and southern italy (vila et al., 2001; amorim et al., 2010; bennouna et al., 2010; laza-martinez et al., 2011; ciminiello et al., 2013) and its morphotype has also been reported along the northern african coast (turki, 2005; turki et al., 2006; mabrouk et al., 2011, 2012). moreover, penna et al. (2012) found a new genotype, probably corresponding to a new species of ostreopsis, in both the atlantic coast (canary islands) and mediterranean sea (greece and cyprus). ostreopsis spp. have been recorded in other temperate areas as well. in japan, toxic strains of both o. siamensis and o. ovata have been recorded not only in sub-tropical southern okinawan waters in late 1970s (fukuyo, 1981; nakajima et al., 1981; yasumoto et al., 1987) but also in the more temperate northern waters of western kyushu, eastern miyazaki, kochi and shikoku (taniyama et al., 2003; adachi et al., 2008; sagara, 2008). ostreopsis cf. siamensis cells were found in temperate new south wales and tasmanian waters (murray, 2010) and, in the latter, in the gut contents of wild mussels (pearce et al., 2001). moreover, o. cf. siamensis has become a major bloom former in new zealand, causing extensive mats covering seaweeds in the eastern northland waters, and has been reported as far south as temperate wellington waters (rhodes et al., 2000, 2010; shears and ross, 2009); in the northern new zealand waters also o. lenticularis and o. cf. ovata have been recorded (chang et al., 2000). finally, ostreopsis cf. ovata and o. cf. siamensis were a constant component of the epiphytic communities during the summer–fall period in peter the great bay, sea of japan since 2006 (selina and orlova, 2010; selina et al., 2014). toxin profile and factors affecting toxin production ostreopsis species produce different toxins, mostly belonging to the palytoxin group. among the nine species of the genus ostreopsis, toxicity has been demonstrated in o. siamensis, o. mascarenensis, o. lenticularis and o. cf. ovata (nakajima et al., 1981; yasumoto et al., 1987; holmes et al., 1988; mercado et al., 1994; meunier et al., 1997; lenoir et al., 2004; ciminiello et al., 2006; scalco et al., 2012; uchida et al., 2013; brissard et al., 2015; garcía-altares et al., 2015). moreover, o. heptagona was determined to be toxic as methanol extracts of culture of this species isolated from knight key (florida) were weakly toxic to mice (babinchak, according to norris et al., 1985). palytoxin (pltx) has a molecular formula of c129h221n3o54 and a molecular weight of 2680 da (moore and bartolini, 1981). it has been primarily isolated from the marine zoanthid palythoa toxica (moore and scheuer, 1971), from which the name comes. pltx is a very complex molecule with both lipophilic and hydrophilic groups and is slightly less toxic than maitotoxin in total potency. the pltx analogues produced by ostreopsis species have a similar chemical structure as the parent pltx, as well as a similar mode and site of action. ostreocin-d was the first pltx analogue isolated from cultures of o. siamensis (usami et al., 1995; ukena et al., 2001). this compound has the chemical formula c127h220n3o53 and a molecular weight of 2634 da, a little lower than pltx. another pltx analogue, mascarenotoxin (mctx), was isolated from o. mascarenensis (lenoir et al., 2004) and o. cf. ovata (rossi et al., 2010; scalco et al., 2012). the molecular weight of the three identified mascarenotoxin congeners ranges from 2500 to 2628 da (rossi et al., 2010). in addition, a third pltx analogue was isolated from o. cf. ovata, the ovatoxin (ovtx) (ciminiello et al., 2008, 2010, 2012a; rossi et al., 2010). mediterranean cultures of o. cf. ovata were found to produce isobaric palytoxin, ovatoxin-a, b, c, d, e, f, g and h and mascarenotoxin-a and c (scalco et al., 2012; garcía-altares et al., 2014; brissard et al., 2015). on the contrary, the mediterranean o. cf. siamensis strain seems to be devoid of any appreciable toxicity (ciminiello et al., 2013). finally, ostreotoxins (produced by o. lenticularis) do not display the same mode and site of action as pltx-analogues and the classification of these compounds as pltx analogues is still unclear (mercado et al., 1994; meunier et al., 1997). as far as the studies on the action mechanism are concerned, almost all studies refer to the commercial pltx standard (tubaro et al., 2014). pltx targets membrane sodium-potassium pumps (na+/k+-atpase) responsible for maintaining ionic gradients (artigas and gadsby, 2003). characteristic aspects of pltx include delayed haemolysis with a loss of potassium, converting na/k pump into a non-specific ionic channel leading to the disruption of ion homeostasis exerted on excitable tissues (habermann et al., 1981). this results in nausea, vomiting, hyper-salivation, abdominal cramps, diarrhoea, numbness of extremities, severe muscular spasms and respiratory distress (yasumoto et al., 1986; alcala et al., 1988; kodama et al., 1989). on the contrary, the possible effects recorded in non-excitable cells are less clear: in this case, the toxin can affect different sets of proteins and signalling pathways, stressing the complexity of the mode of action of pltx (bellocci et al., 2011; rossini and bigiani, 2011; wattenberg, 2011). for example, protein kinases involved in the control of cell proliferation can be activated by pltx (wattenberg, 2011), providing a possible role of this toxin in the tumour-promoting activity (fujiki et al., 1986; fujiki and suganuma, 2009). however, pltx has been also shown to possess a potent cytotoxic activity (bellocci et al., 2011). environmental factors (e.g. temperature, salinity, light, nutrients) and the characteristic of the strains (e.g., isolation 3 no nco mm er cia l u se on ly 4 s. accoroni and c. totti site, growth phase and the age of the strain) have all been shown to affect toxin content in ostreopsis cultures (guerrini et al., 2010; ciminiello et al., 2012a,b; pezzolesi et al., 2012; scalco et al., 2012; vanucci et al., 2012b). several phytoplankton species produce more toxins per cell when growing under suboptimal conditions (johansson and granéli, 1999a,b; etheridge and roesler, 2005). in the same way, some authors observed that o. cf. ovata isolated from both the tyrrhenian and adriatic seas produces higher toxin contents per cell when growing under suboptimal temperature and salinity conditions (granéli et al., 2011; pezzolesi et al., 2012; vidyarathna and granéli, 2013), which differ among strains from different geographical areas (see paragraph 7). on the contrary, optimal nutrient conditions seem to be required for toxin production and both pand n-depleted media decreased o. cf. ovata toxicity (vanucci et al., 2012b). the growth phase affects ostreopsis toxicity as well. although toxin production rate has been found to increase during the exponential phase (pezzolesi et al., 2014), toxins concentration on a per cell basis increased from the exponential to the senescent phase, independently of the growth conditions (guerrini et al., 2010; pistocchi et al., 2011; vanucci et al., 2012a, 2012b). this behaviour was recently explained (pinna et al., 2015) as due to the strong influence of the internal nutrient status (i.e., carbon to nutrient ratio) on toxin synthesis. on the contrary, the toxin profile of o. cf. ovata was relatively stable during the growth stages and independent of culture conditions (pistocchi et al., 2011; scalco et al., 2012; pezzolesi et al., 2014). vectors of exposition and effects on organisms the main vectors for ostreopsis intoxication of humans include marine aerosol (casabianca et al., 2013, 2014; ciminiello et al., 2014), direct contact (tichadou et al., 2010) and the per os ingestion (the latter mainly associated to clupeotoxism syndrome, see below). blooms of o. cf. ovata caused serious problems on human health, mainly due to inhalation of sea water droplets containing ostreopsis cells or fragments and/or aerosolized toxins (gallitelli et al., 2005; kermarec et al., 2008; tichadou et al., 2010; honsell et al., 2011; del favero et al., 2012). one of the most intense episode occurred in summer 2005, when about 200 people exposed to marine aerosols along the ligurian coasts required medical first aid due to similar symptoms of respiratory intoxications, and 20 persons were subjected to extended hospitalization (brescianini et al., 2006; durando et al., 2007). the typical intoxication symptoms of ostreopsis aerosol and direct contact exposure (fever, dyspnoea, broncho-constriction, conjunctivitis and skin irritations) resolve within a few days. regarding the oral ingestion, although ostreopsis has not been confirmed as the source of toxin in clupeotoxism yet, it was strongly suspected of that intoxication in several events (e.g., onuma et al., 1999; randall, 2005). clupeotoxism is one of human intoxications due to consumption of contaminated sardines and herrings (clupeidae) or anchovies (engraulidae). symptomology of clupeotoxism is similar to that of ciguatera (yasumoto et al., 1986), though the former has a much higher mortality rate (onuma et al., 1999). several outbreaks were reported in tropical insular areas of the pacific and the caribbean during the last 30 years (yasumoto et al., 1986; fukui et al., 1987; gleibs et al., 1995). recently, clupeotoxism occurred in the southwestern indian ocean, mainly in madagascar where palytoxin analogues were involved in fatalities occurred after consumption of sardinella fish (yasumoto, 1998; hansen et al., 2001). ostreopsis toxins may contaminate seafood: ostreocin-d produced by o. cf. siamensis were accumulated in wild mussels (mytilus edulis planulatus) from tasmanian coasts (pearce et al., 2001). rhodes et al. (2002), feeding new zealand mussels (perna canaliculus), pacific oysters, and scallops (pecten novaezealandiae) with o. cf. siamensis cells detected trace amounts of palytoxinlike compounds in some of the fed animals. in the mediterranean sea, aligizaki et al. (2008) analysing field samples of shellfish (mytilus galloprovincialis, venus verrucosa, modiolus barbatus) reported that shellfish toxicity coincided with seasonal peaks in ostreopsis abundance, providing the most compelling evidence to date that ostreopsis-borne palytoxin analogues likely accumulate in shellfish. although it has been shown that the oral toxicity of palytoxin and 45-hydroxy palytoxins is about 1000fold less than that observed by intraperitoneal injection (sosa et al., 2009; munday, 2011; tubaro et al., 2011), a regulatory threshold of 30 µg kg–1 has been proposed for shellfish flesh (efsa, 2009). however, the effects of the ingestion of products contaminated by o. cf. ovata toxin are still unknown. ostreopsis blooms are often accompanied by mortality of benthic marine organisms, such as sea urchins, limpets, mussels, crustaceans, holothurians, sponges and even macroalgae (di turi et al., 2003; shears and ross, 2009, 2010; accoroni et al., 2011). in fact, several recent studies have shown that ostreopsis toxicity affects also various marine organisms, both invertebrates and fish (gorbi et al., 2012, 2013; simonini et al., 2011; faimali et al., 2012; pezzolesi et al., 2012; privitera et al., 2012; carella et al., 2015), interfering with embryonic development as well (pagliara and caroppo, 2012). a study carried out on natural banks of mytilus galloprovincialis sampled during various phases of o. cf. ovata bloom in the north-western adriatic sea (mediterranean sea) demonstrated a significant accumulation of algal toxins in mussels exposed, which exhibited a marked inhibino nco mm er cia l u se on ly ostreopsis in temperate areas tion of the na+/k+-atpase activity and alterations of immunological, lysosomal and neurotoxic responses (gorbi et al., 2012). life cycle and cyst formation as the largest part of marine dinoflagellates, ostreopsis has a haplontic life cycle with a dominant motile haploid biflagellate stage (pfiester and anderson, 1987; litaker et al., 2002). the asexual and sexual reproduction of ostreopsis has been studied mainly in ostreopsis cf. ovata, both in the field and in cultures isolated from the mediterranean sea (bravo et al., 2012; accoroni et al., 2014). the asexual reproduction occurs with the division of haploid vegetative cells that takes place in the motile stage: cells divide by desmoschisis in the sagittal plane, with each daughter cell inheriting part of the parental thecal plates (bravo et al., 2012). in general, sexual reproduction of dinoflagellates usually begins with the production of gametes that can fuse with each other forming a diploid zygote. the actual details of gamete formation vary, but in any case, two haploid cells (both typically motile) fuse to yield a diploid cell. the sexual reproduction in o. cf. ovata has been observed in both natural and cultured populations (bravo et al., 2012). gamete pairs are observed in either intercrosses or intracrosses of different strains (bravo et al., 2012) and nutrient limitation seems to stimulate sexual reproduction, which however occurs also in normal culture conditions (accoroni et al., 2014). in o. cf. ovata, two types of mating gametes were identified: i) gametes joined by epitheca, with the point of attachment positioned almost centrally (bravo et al., 2012), and ii) gametes joined laterally with the two cingula perpendicular to each other, with melting of the two thecae (accoroni et al., 2014), as observed in coolia monotis by faust (1992). in dinoflagellates, the newly formed motile diploid cell produced when karyogamy and plasmogamy are complete, is known as a planozygote (pfiester, 1989). the planozygote of some species undergoes meiosis in the plankton, while that of most other dinoflagellates swims for a variable amount of time (hours to weeks), sheds its flagella, rounds up, and settles to form a nonmotile hypnozygote. hypnozygotes often differentiate into long-term resting stages (hypnocysts) that accumulate in sediments and may remain dormant for years before germinating (wall, 1975; anderson et al., 1987). different types of cysts were identified for o. cf. ovata; some of them are non-dormant as germinate within 3 days (temporary cysts), while some others are resting cysts able to germinate for as long as 5 months after their formation. a study on the life cycle of o. cf. ovata conducted with northern adriatic strains highlighted that resting cysts germinated in laboratory conditions at 25°c, but not at 21°c (accoroni et al., 2014). the presence of a temperature threshold affecting cyst germination in laboratory conditions would highlight that temperature could represent a key factor for ostreopsis cf. ovata bloom onset also in natural environments. role of environmental parameters on blooms hydrodynamic conditions several studies considered hydrodynamic condition as the main factor affecting ostreopsis bloom trends, highlighting that higher abundances are observed in sheltered sites compared with exposed ones (barone, 2007; shears and ross, 2009; totti et al., 2010; mabrouk et al., 2011). the abundances of benthic dinoflagellates are highly affected by wave action, since they are only loosely attached to the substrata and can be easily removed and re-suspended in the water column, although in literature we often found controversial data. chang et al. (2000) suggested that o. cf. siamensis was more abundant on the northern new zealand’s eastern coast because it is a less energetic, more stable environment than the western one. this suggestion has been confirmed later by shears and ross (2009) that observed higher ostreopsis siamensis abundances at sheltered vs exposed locations. vila et al. (2001), on the other hand, concluded that ostreopsis sp. prefers ‘moderately shaken’ waters in the nw mediterranean and parsons and preskitt (2007) observed higher abundance of o. cf. ovata on the windward coast of the island of hawaii, whereas ostreopsis sp.1 was more prevalent on the leeward coast. observations in the northern adriatic sea highlighted that 1) significantly higher abundances were observed in the sheltered sites compared with the exposed ones; 2) hydrodynamics may have an important effect on the temporal variability of bloom, because stormy events can result in a sudden decrease of cell abundances on the benthic substrata, with cell proliferation being re-established at high densities after some days of calm sea conditions (totti et al., 2010; accoroni et al., 2012a). moreover, it has been highlighted that turbulence can affect o. cf. ovata growth rate and consequently its cell size (accoroni et al., 2012b). the effect of hydrodynamics has been separately tested for each bloom phase, i.e. initial (no more than 102 cells cm–2 recorded over all substrata), proliferation (period of intense cell division, when o. cf. ovata rapidly increased abundances until reaching maximum peak) and decline phase (decrease in cell abundances and bloom decline), and it has been shown that cell abundances in sheltered sites were significantly higher than those in the 5 no nco mm er cia l u se on ly 6 s. accoroni and c. totti exposed ones during the proliferation phase (accoroni et al., 2012b). this result suggested that hydrodynamics affect o. cf. ovata abundances mainly during phases when the highest abundances are reached. in fact, a well-developed benthic mat (i.e., brownish pellicle loosely attached to benthic substrata) is produced only during the most intense proliferation period (totti et al., 2010) and this structure is easily removed by effect of the hydrodynamic conditions. this result may explain why such effects linked to hydrodynamic conditions were not observed in those areas where high abundances were not reached and a mat did not develop, as observed in the tyrrhenian sea (zingone, personal communication), and in johnston atoll (pacific ocean) (richlen and lobel, 2011). water temperature many authors suggested that ostreopsis spp. need relatively high temperatures to proliferate, proposing that the global warming might have influenced ostreopsis expansion in temperate areas such as the mediterranean sea (hallegraeff, 2010; granéli et al., 2011), but a more careful analysis of literature data shows that temperature role is not the same in all coastal areas around the world (tab. 1). ostreopsis blooms are summer events in temperate areas, although comparing the bloom trend in several mediterranean areas it can be observed that peaks can occur from spring to autumn, with a certain inter-annual variability. in the northern adriatic sea, the peaks of the blooms occur generally in september-october (monti et al., 2007; totti et al., 2010; accoroni et al., 2015a). on the contrary, in the ligurian sea highest cell abundances were mostly recorded in mid-summer (end of july) (mangialajo et al., 2008). vila et al. (2001) observed abundance peaks for ostreopsis sp. in the north-western mediterranean even early, in springtime. also in the aegean sea, spatharis et al. (2009) found o. cf. ovata being most abundant in may, contrarily to what observed in the same area by aligizaki and nikolaidis (2006) that reported peak abundances from midsummer to late fall. experimental studies about the response of ostreopsis growth to temperature provide a possible explanation for the above different temporal trends observed in situ: guerrini et al. (2010) and pezzolesi et al. (2012) highlighted that o. cf. ovata strains from different italian coasts (in the tyrrhenian and adriatic seas) displayed different growth temperature optima that parallel with the in situ temperature values typical of the blooming period of the single strain. scalco et al. (2012) observed that mediterranean o. cf. ovata strains grew within a window of 1830°c with the best performances recorded between 22 and 26°c, suggesting that o. cf. ovata is adapted to intermediate temperatures and day length conditions, such are those recorded in the natural environment at the beginning of summer and/or of autumn (mangialajo et al., 2011). in ta b. 1 .s um m ar y of e nv ir on m en ta l c on di tio ns r ec or de d du ri ng o st re op si s sp ec ie s bl oo m s in te m pe ra te a re as . v al ue s re fe r to p er io ds o f o st re op si s m ax im um p ro lif er at io n; d at a no t ex pl ic at ed in th e m ai n te xt o f t he re fe re nc es , w he re ve r p os si bl e, h av e be en e xt ra po la te d fr om ta bl es o r f ig ur es . a re a s pe ci es s am pl ed s ub st ra ta m ax im um te m pe ra tu re (° c ) s al in it y n ut ri en ts (µ m ) r ef er en ce p ro lif er at io n pe ri od m ed ite rr an ea n se a, o st re op si s cf . o va ta m ac ro al ga e, in ve rt eb ra te s, se pt em be ro ct ob er 1 6. 827 .9 3 1. 339 .3 0 .5 519 .4 (d in ) m on ti et a l., 2 00 7; t ot ti et a l., 2 01 0; n a dr ia tic s ea r oc ks , w at er c ol um n 0 .0 10. 49 (p o 4) a cc or on i e t a l., 2 01 1, 2 01 2a , 2 01 5a ; m an gi al aj o et a l., 2 01 1 m ed ite rr an ea n se a, o st re op si s cf . o va ta w at er c ol um n a ug us t 2 6. 427 .1 3 6. 838 .4 1 0. 337 .1 0 (t n ) u ng ar o et a l., 2 00 5 s a dr ia tic s ea 0 .1 10. 15 (p o 4) m ed ite rr an ea n se a, o st re op si s cf . o va ta w at er c ol um n a ug us tea rl y o ct ob er 2 4. 528 to gn et to e t a l., 1 99 5 ty rr he ni an s ea n m ed ite rr an ea n se a, o st re op si s cf . o va ta m ac ro al ga e, w at er c ol um n ju ly -a ug us t 2 2. 630 3 8. 038 .2 c im in ie llo e t a l., 2 00 6; l ig ur ia n se a m an gi al aj o et a l., 2 00 8, 2 01 1; c oh u et a l., 2 01 1 n w m ed ite rr an ea n se a, o st re op si s cf . o va ta m ac ro al ga e, w at er c ol um n ju ly -a ug us t 19 .2 -2 1. 5 m an gi al aj o et a l., 2 01 1 g ul f o f l io n n w m ed ite rr an ea n se a, o st re op si s cf . o va ta m ac ro al ga e, s of t s ed im en ts , e nd m ar ch o ct ob er 1 828 .3 3 0. 338 .1 0 .7 67. 74 (d in ) v ila e t a l., 2 00 1; c at al an s ea o st re op si s cf . s ia m en si s w at er c ol um n 0 .1 10. 86 (p o 4) m an gi al aj o et a l., 2 01 1; c ar ni ce r e t a l., 2 01 5 m ed ite rr an ea n se a, o st re op si s cf . o va ta m ac ro al ga e, a ng io sp er m s, m ay -e ar ly n ov em be r 1 3. 929 .7 085 (d in ) a lig iz ak i a nd n ik ol ai di s, 2 00 6; a eg ea n se a o st re op si s cf . s ia m en si s s of t s ed im en ts , w at er c ol um n 0 .5 -6 .5 (d ip ) s pa th ar is e t a l., 2 00 9 m ed ite rr an ea n se a, o st re op si s cf . s ia m en si s a ng io sp er m s a ug us t o ct ob er 20 -2 7 3 6. 637 t ur ki , 2 00 5 g ul f o f t un is n ew z ea la nd o st re op si s ov at a m ac ro al ga e, w at er c ol um n e nd f eb ru ar ya pr il 1 7. 822 .1 1 .6 -3 .8 (d in ) c ha ng e t a l., 2 00 0; o st re op si s si am en si s 0 .3 31. 10 (p o 4) s he ar s an d r os s, 2 00 9 o st re op si s le nt ic ul ar is se a of j ap an o st re op si s cf . o va ta m ac ro al ga e a ug us to ct ob er 9 -2 5 30 -3 4 s el in a an d o rl ov a, 2 01 0; o st re op si s cf . s ia m en si s se lin a et a l., 2 01 4 no nco mm er cia l u se on ly ostreopsis in temperate areas the same way, tawong et al. (2015) observed that optimal and tolerable temperature conditions differ among ostreopsis cf. ovata subclades: strains of o. cf. ovata thailand subclade and o. cf. ovata south china sea subclade showed the semi-optimal temperature ranges of 22.727.4°c and 27.9-30.8°c, with optimal temperature of 25°c and 30°c, respectively. although the reaching of the highest abundances of ostreopsis is not in concomitance with the highest water temperature values in all areas, a temperature threshold would seem to be important to let the bloom start: a study conducted along the conero riviera (northern adriatic sea) showed that, although the bloom peak occurred in late summer (when temperatures ranged between 18.8 and 24°c, decreasing from the seasonal maximum), the bloom onset was always observed at higher temperature (25-28.6°c), suggesting that ostreopsis needs to reach a fairly well fixed temperature threshold to start its bloom, probably in relation to the cyst germination that generally occurs at around 25°c (accoroni et al., 2014). as in this area the bloom can persist until temperature values are much lower (14.417.5°c) than that threshold, one would guess that once ostreopsis cysts are germinated, its vegetative forms seem to actively proliferate even if temperature values decrease. this discrepancy between optimal temperature range for the cysts germination and the algal growth has been reported for the dinoflagellate scrippsiella trochoidea as well by binder and anderson (1987) in experimental conditions. however, in the northern adriatic sea the bloom onset is often observed about 30 days after the reaching of the 25 °c-temperature threshold, suggesting that other environmental factors, besides temperature, may affect the development of o. cf. ovata blooms. in this regard, o. cf. ovata blooms appear to be triggered by a combination of optimal temperature and available nutrients, where the temperature threshold plays a key role on the germination of o. cf. ovata cysts and an n:p ratio around the redfield value is a necessary condition to allow cell proliferation. role of salinity and nutrients contrasting results are reported in the literature about the effect of salinity on the development of ostreopsis blooms. rhodes et al. (2000) observed that o. cf. siamensis had a preferred salinity range of 28-34 in cultures isolated from northland (new zealand). salinity measured in the north-western mediterranean sea during the ostreopsis spp. blooms showed values around 37-38 (vila et al., 2001; mangialajo et al., 2008) and similar salinity levels were measured in the gulf of trieste (northern adriatic sea) by monti et al. (2007) and in the southern adriatic sea by ungaro et al. (2005); in the conero riviera (northern adriatic sea) during the ostreopsis blooms occurred from 2007 to 2012, salinity ranged in a much wider range, from 31.3 to 39.3 (accoroni et al., 2015a). several authors suggested that benthic dinoflagellates proliferation is favoured by low salinity waters. in the virgin islands, abundance maxima of ostreopsis were correlated with the period of maximal rainfall (carlson and tindall, 1985), and the same negative correlation with salinity has been found for o. ovata along the hawaiian coasts (parsons and preskitt, 2007). on the contrary, in the north-western coast of cuba, in catalonian and in french mediterranean coasts, delgado et al. (2006), blanfuné et al., (2015) and carnicer et al., (2015) suggested that low salinity values possibly hinder ostreopsis spp. as they found conspicuously lower ostreopsis abundances in sites more affected by river plumes (i.e., jaimanitas river, rhone river and ebro delta) than in the rest of the studied areas. no significant correlation between cell abundances and salinity values were found in the gulf of mexico for o. heptagona (okolodkov et al., 2007) and in the conero riviera for o. cf. ovata (accoroni et al., 2015a) when the effect of salinity has been investigated in experimental conditions, it has been shown that such effect may be strain-specific: tawong et al. (2015) showed that optimal and tolerable salinity conditions differed among ostreopsis cf. ovata subclades as the optimal salinities for the o. cf. ovata thailand and south china sea subclades were 30 and 25, respectively. pezzolesi et al. (2012) demonstrated that an adriatic o. cf. ovata strain cultured at salinity ranging from 26 to 40 showed that although the lowest growth rates were observed at the lowest salinity, growth rates did not significantly differ in different salinity conditions. indeed, the relationships between algal blooms in the field and salinity are more complicated, and other factors, such as nutrient levels (which are typically associated to low salinity waters) have to be considered. recent studies have provided increasing evidence of a link between the nutrient enrichment of coastal waters (anthropogenic eutrophication) and harmful algal events (glibert and burkholder, 2006; glibert et al., 2010). however, there is very limited information on the relationships between nutrient concentrations and the occurrence of ostreopsis blooms (tab. 1). vila et al. (2001) and cohu et al. (2011) in northwestern mediterranean sea and shears and ross (2009) in north-eastern new zealand did not find any relation between epiphytic o. cf. ovata abundance and nutrients, while parsons and preskitt (2007) found that ostreopsis sp.1 abundance was positively correlated with nutrient concentrations in the waters surrounding hawaii. in the conero riviera (northern adriatic sea), although no clear relationship was found between nutrient concentrations and o. cf. ovata abundances, it was observed that in the bloom onset period, po4 concentrations were significantly higher than in the rest of the study period. interestingly, the following bloom development is maintained in a con7 no nco mm er cia l u se on ly 8 s. accoroni and c. totti dition of elevated n:p ratios, suggesting that such blooms may be initiated at low n:p levels (possibly stimulated by a ‘flush’ of nutrients or organic materials) that may allow the newly germinating cells to increase growth rate while other adaptive mechanisms (e.g., metabolic dissipatory strategies, allelopathic and mixotrophic interactions), would enable the maintenance of blooms at less than maximal growth rates and at not-optimal n:p ratios (accoroni et al., 2015a). a decrease in n:p ratio values has previously been associated with the onset of a number of planktonic dinoflagellate blooms as well (hodgkiss and ho, 1997; zhang and hu, 2011; glibert et al., 2012). in this regard, experimental studies conducted on different adriatic o. cf. ovata strains showed that the depletion of p was proportionately more rapid than that of n, highlighting the strong p demand of this dinoflagellate (vanucci et al., 2012b; pezzolesi et al., 2014). moreover, further studies are needed to clarify the trophic behaviour of ostreopsis spp., considering that for these species also mixotrophy was hypothesized (barone, 2007; burkholder et al., 2008), which may play an important role in ostreopsis development, as already observed in other potentially toxic microalgae (cucchiari et al., 2008; heisler et al., 2008). depth the role of depth on ostreopsis abundances along the conero riviera (northern adriatic sea) was investigated in 2007 (totti et al., 2010) in target sites where samples were collected at depths comprised between 0.5 and 10 m. o. cf. ovata abundances showed a significant decrease with depth, in agreement with what observed by richlen and lobel (2011) and cohu and lemée (2012), suggesting a potential effect of light intensity. this may explain why ostreopsis blooms mainly develop in shallow waters. however, such effect has not been observed in shallow sites affected by high hydrodynamics, such as on the fringing reefs of the higher infralittoral shelf, where o. cf. ovata abundances were lower than those recorded immediately deeper, due to the hydrodynamic effect of wave actions (totti et al., 2010). as previously suggested, the role of depth may be related to light availability, although there are conflicting data about the relationship between light intensity and ostreopsis growth in experimental conditions. morton et al. (1992) reported that o. cf. siamensis and o. heptagona isolated from the florida keys displayed maximal growth at approximately 200 µmol photons m–2 s–1 and did not grow rapidly at over 240 µmol photons m–2 s–1 (i.e., >10% full sun light). scalco et al. (2012), analysing the growth performance of some italian o. cf ovata strains, observed that this species grew better at relatively low photon flux density (50 instead of 200 µmol photons m–2 s–1). yamaguchi et al. (2014) observed that ostreopsis sp. from japan grew proportionally when light intensity was increased from 49.5 to 199 µmol photons m–2 s–1, but its growth appeared to be inhibited slightly at >263 µmol photons m–2 s–1. heil et al. (1993) observed that o. cf. siamensis cultured in spinner flasks would stay planktonic when light intensities are maintained low (25 µmol photons m–2 s–1), but would produce mucus and settle at the bottom at higher light intensities (75 µmol photons m–2 s– 1). this has been interpreted as a protective measure to shade the cells. therefore, in experimental conditions ostreopsis sp. seems to suffer too high light intensities, therefore being exposed to potentially detrimental consequences of photodamaging. in field conditions, there are only few data about the relationship between ostreopsis abundances and light intensity, which, moreover, seem to disagree with the experimental evidences described above. a study carried out along several italian coastal areas affected by ostreopsis blooms showed that during the blooms, the values of light intensity at depth of ostreopsis sampling were quite high, up to 1800 µmol photons m–2 s–1 (ispra, 2012). however, it is known that light availability amidst macroalgal vegetation is generally low (raniello et al., 2004), and ballantine et al. (1988) suggested that ostreopsis cells can migrate to shaded areas of the algal host thallus to escape high light levels. anyway, further studies are required to clarify the real role of light intensity in the bloom dynamics of ostreopsis. substratum ostreopsis has often been indicated to be preferentially epiphytic on macroalgae (bomber et al., 1989; vila et al., 2001), although it has been recorded on a variety of other substrata (table 1), including marine angiosperms (turki, 2005; aligizaki and nikolaidis, 2006; turki et al., 2006; battocchi et al., 2010; mabrouk et al., 2012), rocks (bottalico et al., 2002; totti et al., 2010; accoroni et al., 2011), coral rubble (norris et al., 1985), soft sediments (vila et al., 2001; aligizaki and nikolaidis, 2006), and invertebrates (bianco et al., 2007; totti et al., 2010). they also can be found as free-living in the plankton (faust and morton, 1995; tognetto et al., 1995; chang et al., 2000; totti et al., 2010; accoroni et al., 2011). the possibility of ostreopsis to colonize a variety of substrata, living either as epiphytic, epilithic, or epizoic, indicates that this species is not an obligate epiphyte. a number of studies underlined the importance of host thallus architecture (lobel et al., 1988; bomber et al., 1989). vila et al. (2001) observed that three-dimensional flexible thalli are more suitable for the growth of ostreopsis spp. it has been suggested that the higher abundances found in branched than in flattened thalli, might be explained by a different response of such thallus morphotypes to the wave action (totti et al., 2010). no nco mm er cia l u se on ly ostreopsis in temperate areas indeed, the relationships between ostreopsis and macrophytes are more complicated. in studies carried out on natural populations of o. cf. ovata, significantly higher abundances were reported on pebbles than on macroalgae (totti et al., 2010; accoroni et al., 2011), suggesting that living substrata allow lower concentration of epibionts than any other substrate, probably due to the production of some hypothetical allelopathic compounds (jin and dong, 2003). in this regard, a study conducted to assess any possible allelopathic interactions between ostreopsis cf. ovata and macroalgae showed that all the investigated seaweeds [dictyota dichotoma (brown alga), rhodymenia pseudopalmata (red alga) and ulva rigida (green alga)] exerted negative effects toward the benthic dinoflagellate, with the highest inhibitory effect observed in d. dichotoma and the lowest in r. pseudopalmata (accoroni et al., 2015b). unresolved issues despite the number of studies on ostreopsis biology, ecology and toxin production and actions, several aspects about the environmental concerns associated with this genus remain still unclear. regarding the action mechanism of the implicated toxins, almost all studies refer to the commercial pltx standard, while a more accurate analysis should be addressed on both effects and action mechanism of all toxins produced by ostreopsis species. moreover, among the vectors of ostreopsis intoxication i.e., marine aerosol, direct contact and per os ingestion, the latter needs certainly further studies given its possible implications on human health almost unknown in temperate areas nowadays. numerous studies have highlighted the influence of environmental factors on bloom dynamics of ostreopsis and the complexity of conditions leading to blooms of this dinoflagellate is becoming clearer but not totally understood. although the mechanisms for bloom onset seems clarified, those driving both bloom development and decline are still far from being understood, and other both abiotic and biotic factors, such as the interactions with other organisms and the ability to use organic forms of nutrients, should be investigated. actually, several hab genera have been shown to use organic forms of nutrients for their nutritional demands (cucchiari et al., 2008; heisler et al., 2008) and a mixotrophic behavior has been hypothesized in ostreopsis (barone, 2007; burkholder et al., 2008; pinna et al., 2015). effects of biotic interactions on ostreopsis should be considered as well and only few studies have been carried out on bacteria (vanucci et al., 2012a), diatoms (pichierri et al 2015) and macroalgae (accoroni et al., 2015b). in this regard, the role of viruses, bacteria and parasites in both cysts formation and bloom termination have been recognized in several microalgae (nagasaki et al., 1994; nagasaki et al., 2000; tarutani et al., 2001; mizumoto et al., 2008; garcés et al., 2013) and should be studied in ostreopsis bloom dynamics as well. acknowledgments this study was partially supported by ispra-italian ministry of the environment, murst (prin 2007), and enpi cbcmed m3-habs project. references accoroni s, romagnoli t, colombo f, pennesi c, di camillo cg, marini m, battocchi c, ciminiello p, dell’aversano c, dello iacovo e, fattorusso e, tartaglione l, penna a, totti c, 2011. ostreopsis cf. ovata bloom in the northern adriatic sea during summer 2009: ecology, molecular characterization and toxin profile. mar. pollut. bull. 62:2512-2519. accoroni s, colombo f, pichierri s, romagnoli t, marini m, battocchi c, penna a, totti c, 2012a. ecology of ostreopsis cf. ovata blooms in the northwestern adriatic sea. cryptogam. algol. 33:191-198. accoroni s, romagnoli t, pichierri s, colombo f, totti c, 2012b. morphometric analysis of ostreopsis cf. ovata cells in relation to environmental conditions and bloom phases. harmful algae 19:15-22. accoroni s, romagnoli t, pichierri s, totti c, 2014. new insights on the life cycle stages of the toxic benthic dinoflagellate ostreopsis cf. ovata. harmful algae 34:7-16. accoroni s, glibert pm, pichierri s, romagnoli t, marini m, totti c, 2015a. a conceptual model of annual ostreopsis cf. ovata blooms in the northern adriatic sea based on the synergic effects of hydrodynamics, temperature, and the n:p ratio of water column nutrients. harmful algae 45:14-25. accoroni s, percopo i, cerino f, romagnoli t, pichierri s, perrone c, totti c, 2015b. allelopathic interactions between the hab dinoflagellate ostreopsis cf. ovata and macroalgae. harmful algae 49:147-155. adachi m, ikegami t, yamaguchi h, 2008. the genus ostreopsis in the coastal regions in tosa bay, southern part of japan, p. 1. in: proc. 13th int. conf. on harmful algae, hong kong. alcala ac, alcala lc, garth js, yasumura d, yasumoto t, 1988. human fatality due to ingestion of the crab demania reynaudii that contained a palytoxin-like toxin. toxicon 26:105-107. aligizaki k, nikolaidis g, 2006. the presence of the potentially toxic genera ostreopsis and coolia (dinophyceae) in the north aegean sea, greece. harmful algae 5:717-730. aligizaki k, katikou p, nikolaidis g, panou a, 2008. first episode of shellfish contamination by palytoxin-like compounds from ostreopsis species (aegean sea, greece). toxicon 51:418-427. amorim a, veloso v, penna a, 2010. first detection of ostreopsis cf. siamensis in portuguese coastal waters. harmful algae news 42:6-7. anderson dm, taylor cd, armbrust ev, 1987. the effects of darkness and anaerobiosis on dinoflagellate cyst germination. limnol. oceanogr. 32:340-351. 9 no nco mm er cia l u se on ly 10 s. accoroni and c. totti artigas p, gadsby dc, 2003. na+/k+-pump ligands modulate gating of palytoxin-induced ion channels. p. natl. acad. sci. usa 100:501-505. ballantine dl, bardales at, tosteson tr, durst hd, 1985. seasonal abundance of gambierdiscus toxicus and ostreopsis sp. in coastal waters of southwest puerto rico, p. 417-422. in: proc. 5th int. coral reef congr., tahiti. ballantine dl, tosteson tr, bardales at, 1988. population dynamics and toxicity of natural populations of benthic dinoflagellates in southwestern puerto rico. j. exp. mar. biol. ecol. 119:201-212. barone r, 2007. behavioural trait of ostreopsis ovata (dinophyceae) in mediterranean rock pools: the spider’s strategy. harmful algae news 33:1-3. battocchi c, totti c, vila m, masó m, capellacci s, accoroni s, reñé a, scardi m, penna a, 2010. monitoring toxic microalgae ostreopsis (dinoflagellate) species in coastal waters of the mediterranean sea using molecular pcr-based assay combined with light microscopy. mar. pollut. bull. 60:1074-1084. bellocci m, sala gl, prandi s, 2011. the cytolytic and cytotoxic activities of palytoxin. toxicon 57:449-459. bennouna a, el attar j, abouabdellah r, palma s, penna a, moita t, 2010. first records of ostreopsis cf. siamensis in moroccan atlantic upwelling waters. harmful algae news 42:1-3. bianco i, sangiorgi v, penna a, guerrini f, pistocchi r, zaottini e, congestri r, 2007. ostreopsis ovata in benthic aggregates along the latium coast (middle tyrrhenian sea), p. 29 (abstract). in: proc. int. symp. on algal toxins, trieste, italy. binder bj, anderson dm, 1987. physiological and environmental control of germination in scrippsiella trochoidea (dinophyceae) resting cysts. j. phycol. 23:99-107. blanfuné a, boudouresque c, grossel h, thibaut t, 2015. distribution and abundance of ostreopsis spp. and associated species (dinophyceae) in the northwestern mediterranean: the region and the macroalgal substrate matter environ. sci. pollut. res.22:12332-12346. bomber jw, aikman ke, 1989. the ciguatera dinoflagellates. biol. oceanogr. 6:291-311. bomber jw, rubio mg, norris dr, 1989. epiphytism of dinoflagellates associated with the disease ciguatera: substrate specificity and nutrition. phycologia 28:360-368. bottalico a, micella p, feliciti gp, 2002. [fioritura di ostreopsis sp. (dinophyta) nel porto di otranto].[article in italian]. in: proc. riunione scientifica annuale del gruppo di lavoro per l’algologia società botanica italiana, chioggia, italy. bravo i, vila m, casabianca s, rodriguez f, rial p, riobó p, penna a, 2012. life cycle stages of the benthic palytoxinproducing dinoflagellate ostreopsis cf. ovata (dinophyceae). harmful algae 18:24-34. brescianini c, grillo c, melchiorre n, bertolotto r, ferrari a, vivaldi b, icardi g, gramaccioni l, funari e, scardala s, 2006. ostreopsis ovata algal blooms affecting human health in genova, italy, 2005 and 2006. euro surveill. 11:e060907.3. brissard c, hervé f, sibat m, séchet v, hess p, amzil z, herrenknecht c, 2015. characterization of ovatoxin-h, a new ovatoxin analog, and evaluation of chromatographic columns for ovatoxin analysis and purification. j. chromatogr. a 1388:87-101. burkholder jm, glibert pm, skelton hm, 2008. mixotrophy, a major mode of nutrition for harmful algal species in eutrophic waters. harmful algae 8:77-93. carella f, sardo a, mangoni o, di cioccio d, urciuolo g, de vico g, zingone a, 2015. quantitative histopathology of the mediterranean mussel (mytilus galloprovincialis l.) exposed to the harmful dinoflagellate ostreopsis cf. ovata. j. invertebr. pathol. 127:130-140. carlson rd, tindall dr, 1985. distribution and periodicity of toxic dinoflagellates in the virgin islands, p. 171-176. in: d.m. anderson, a.w. white and d.g. baden (eds.), toxic dinoflagellates. elsevier. carnicer o, guallar c, andree kb, diogène j, fernández-tejedor m, 2015. ostreopsis cf. ovata dynamics in the nw mediterranean sea in relation to biotic and abiotic factors. environ. res. 143(pt b):89-99. casabianca s, casabianca a, riobó p, franco jm, vila m, penna a, 2013. quantification of the toxic dinoflagellate ostreopsis spp. by qpcr assay in marine aerosol. environ. sci. technol. 47:3788-3795. casabianca s, perini f, casabianca a, battocchi c, giussani v, chiantore m, penna a, 2014. monitoring toxic ostreopsis cf. ovata in recreational waters using a qpcr based assay. mar. pollut. bull. 88:102-109. chang fh, shimizu y, hay b, stewart r, mackay g, tasker r, 2000. three recently recorded ostreopsis spp. (dinophyceae) in new zealand: temporal and regional distribution in the upper north island from 1995 to 1997. new zeal. j. mar. fresh. 34:29-39. ciminiello p, dell’aversano c, dello iacovo e, fattorusso e, forino m, grauso l, tartaglione l, guerrini f, pezzolesi l, pistocchi r, vanucci s, 2012a. isolation and structure elucidation of ovatoxin-a, the major toxin produced by ostreopsis ovata. j. am. chem. soc. 134:1869-1875. ciminiello p, dell’aversano c, dello iacovo e, fattorusso e, forino m, tartaglione l, battocchi c, crinelli r, carloni e, magnani m, penna a, 2012b. unique toxin profile of a mediterranean ostreopsis cf. ovata strain: hr lc-msn characterization of ovatoxin-f, a new palytoxin congener. chem. res. toxicol. 25:1243-1252. ciminiello p, dell’aversano c, dello iacovo e, fattorusso e, forino m, tartaglione l, benedettini g, onorari m, serena f, battocchi c, casabianca s, penna a, 2014. first finding of ostreopsis cf. ovata toxins in marine aerosols. environ. sci. technol. 48:3532-3540. ciminiello p, dell’aversano c, fattorusso e, forino m, magno gs, tartaglione l, grillo c, melchiorre n, 2006. the genoa 2005 outbreak. determination of putative palytoxin in mediterranean ostreopsis ovata by a new liquid chromatography tandem mass spectrometry method. anal. chem. 78:6153-6159. ciminiello p, dell’aversano c, fattorusso e, forino m, tartaglione l, grillo c, melchiorre n, 2008. putative palytoxin and its new analogue, ovatoxin-a, in ostreopsis ovata collected along the ligurian coasts during the 2006 toxic outbreak. j. am. soc. mass spectrom. 19:111-120. ciminiello p, dell’aversano c, dello iacovo e, fattorusso e, forino m, tartaglione l, yasumoto t, battocchi c, giacobbe m, amorim a, penna a, 2013. investigation of toxin profile of mediterranean and atlantic strains of ostreopsis cf. siamensis (dinophyceae) by liquid chromatography-high no nco mm er cia l u se on ly ostreopsis in temperate areas resolution mass spectrometry. harmful algae 23:19-27. cohu s, lemée r, 2012. vertical distribution of the toxic epibenthic dinoflagellates ostreopsis cf. ovata, prorocentrum lima and coolia monotis in the nw mediterranean sea. cah. biol. mar. 53:373-380. cohu s, thibaut t, mangialajo l, labat j-p, passafiume o, blanfuné a, simon n, cottalorda j-m, lemée r, 2011. occurrence of the toxic dinoflagellate ostreopsis cf. ovata in relation with environmental factors in monaco (nw mediterranean). mar. pollut. bull. 62:2681-2691. cucchiari e, guerrini f, penna a, totti c, pistocchi r, 2008. effect of salinity, temperature, organic and inorganic nutrients on growth of cultured fibrocapsa japonica (raphidophyceae) from the northern adriatic sea. harmful algae 7:405-414. del favero g, sosa s, pelin m, d’orlando e, florio c, lorenzon p, poli m, tubaro a, 2012. sanitary problems related to the presence of ostreopsis spp. in the mediterranean sea: a multidisciplinary scientific approach. ann. ist. super. sanità 48:407-414. delgado g, lechuga-deveze ch, popowski g, troccoli l, salinas ca, 2006. epiphytic dinoflagellates associated with ciguatera in the northwestern coast of cuba. rev. biol. trop. 54:299-310. di turi l, lo caputo s, marzano mc, pastorelli am, pompei m, rositani l, ungaro n, 2003. [sulla presenza di ostreopsidaceae (dinophyceae) lungo il litorale barese].[article in italian]. biol. mar. mediterr. 10(2):675–678. durando p, ansaldi f, oreste p, moscatelli p, gasparini lm, icardi g, collaborative group for the ligurian syndromic algal surveillance, 2007. ostreopsis ovata and human health: epidemiological and clinical features of respiratory syndrome outbreaks from a two-year sindromi surveillance, 2005-06, in north-west italy. euro surveill. 12:e070607.1 erratum in euro surveill. 2007;12:following e070705.4. efsa, 2009. marine biotoxins in shellfish palytoxin group. opinion of the scientific panel on contaminants in the food chain. efsa j. 7:1393–1430. etheridge sm, roesler cs, 2005. effects of temperature, irradiance, and salinity on photosynthesis, growth rates, total toxicity, and toxin composition for alexandrium fundyense isolates from the gulf of maine and bay of fundy. deepsea. res. pt. ii 52:2491-2500. faimali m, giussani v, piazza v, garaventa f, corrà c, asnaghi v, privitera d, gallus l, cattaneo-vietti r, mangialajo l, chiantore m, 2012. toxic effects of harmful benthic dinoflagellate ostreopsis ovata on invertebrate and vertebrate marine organisms. mar. environ. res. 76:97-107. faust ma, 1992. observations on the morphology and sexual reproduction of coolia monotis (dinophyceae). j. phycol. 28:94-104. faust ma, 1999. three new ostreopsis species (dinophyceae): o. marinus sp. nov., o. belizeanus sp. nov., and o. carribeanus sp. nov. phycologia 38:92-99. faust ma, morton sl, 1995. morphology and ecology of the marine dinoflagellate ostreopsis labens sp. nov. (dinophyceae). j. phycol. 31:456-463. fujiki h, suganuma m, 2009. carcinogenic aspects of protein phosphatase 1 and 2a inhibitors, p. 221-254. in: n. fusetani and w. kem (eds.), marine toxins as research tools. springer. fujiki h, suganuma m, nakayasu m, hakii h, horiuchi t, takayama s, sugimura t, 1986. palytoxin is a non-12-otetradecanoylphorbol-13-acetate type tumor promoter in twostage mouse skin carcinogenesis. carcinogenesis 7:707-710. fukui m, murata m, inoue a, gawel m, yasumoto t, 1987. occurrence of palytoxin in the trigger fish melichtys vidua. toxicon 25:1121-1124. fukuyo y, 1981. taxonomical study on benthic dinoflagellates collected in coral reefs. bull. japan. soc. sci. fish. 47: 967-978. gallitelli m, ungaro n, addante lm, silver ng, sabba c, 2005. respiratory illness as a reaction to tropical algal blooms occurring in a temperate climate. jama-j. am. med. assoc. 293:2599-2600. garcés e, alacid e, bravo i, fraga s, figueroa ri, 2013. parvilucifera sinerae (alveolata, myzozoa) is a generalist parasitoid of dinoflagellates. protist 164:245-260. garcía-altares m, tartaglione l, dell’aversano c, carnicer o, de la iglesia p, forino m, diogène j, ciminiello p, 2015. the novel ovatoxin-g and isobaric palytoxin (so far referred to as putative palytoxin) from ostreopsis cf. ovata (nw mediterranean sea): structural insights by lc-high resolution msn. anal. bioanal. chem. 407:1191–1204. gleibs s, mebs d, werding b, 1995. studies on the origin and distribution of palytoxin in a caribbean coral reef. toxicon 33:1531-1537. glibert pm, burkholder jm, 2006. the complex relationships between increases in fertilization of the earth, coastal eutrophication and proliferation of harmful algal blooms, p. 341-354. in: e. granéli and j.t. turner (eds.), ecology of harmful algae. springer. glibert pm, allen ji, bouwman af, brown cw, flynn kj, lewitus aj, madden cj, 2010. modeling of habs and eutrophication status, advances, challenges. j. mar. syst. 83:262-275. glibert pm, burkholder jm, kana tm, 2012. recent insights about relationships between nutrient availability, forms, and stoichiometry, and the distribution, ecophysiology, and food web effects of pelagic and benthic prorocentrum species. harmful algae 14:231-259. gorbi s, bocchetti r, binelli a, bacchiocchi s, orletti r, nanetti l, raffaelli f, vignini a, accoroni s, totti c, regoli f, 2012. biological effects of palytoxin-like compounds from ostreopsis cf. ovata: a multibiomarkers approach with mussels mytilus galloprovincialis. chemosphere 89:623-632. gorbi s, avio gc, benedetti m, totti c, accoroni s, pichierri s, bacchiocchi s, orletti r, graziosi t, regoli f, 2013. effects of harmful dinoflagellate ostreopsis cf. ovata exposure on immunological, histological and oxidative responses of mussels mytilus galloprovincialis. fish shellfish immunol. 35:941-950. granéli e, vidyarathna nk, funari e, cumaranatunga prt, scenati r, 2011. can increases in temperature stimulate blooms of the toxic benthic dinoflagellate ostreopsis ovata? harmful algae 10:165-172. guerrini f, pezzolesi l, feller a, riccardi m, ciminiello p, dell’aversano c, tartaglione l, dello iacovo e, fattorusso e, forino m, pistocchi r, 2010. comparative growth and toxin profile of cultured ostreopsis ovata from the tyrrhenian and adriatic seas. toxicon 55:211-220. 11 no nco mm er cia l u se on ly 12 s. accoroni and c. totti habermann e, ahnert-hilger g, chhatwal gs, beress l, 1981. delayed haemolytic action of palytoxin. general characteristics. biochim. biophys. acta 649:481-486. hallegraeff gm, 2010. ocean climate change, phytoplankton community responses, and harmful algal blooms: a formidable predictive challenge. j. phycol. 46:220-235. hansen g, turquet j, quod jp, ten-hage l, lugomela c, kyewalyanga m, hurbungs m, wawiye p, ogongo b, tunje s, rakotoarinjanahary h, 2001. potentially harmful algae of the western indian ocean: a guide based on a preliminary survey. i.o.c., manuals and guides 41. intergovernmental oceanographic commission of unesco. heil ca, maranda l, shimizu y, 1993. mucus-associated dinoflagellates: large scale culturing and estimation of growth rate, p. 501-506. in: t.j. smayda and h. shimizu (eds), toxic phytoplankton blooms in the sea. elsevier. heisler j, glibert pm, burkholder jm, anderson dm, cochlan w, dennison wc, dortch q, gobler cj, heil ca, humphries e, lewitus a, magnien r, marshall hg, sellner k, stockwell da, stoecker dk, suddleson m, 2008. eutrophication and harmful algal blooms: a scientific consensus. harmful algae 8:3-13. hodgkiss ij, ho kc, 1997. are changes in n:p ratios in coastal waters the key to increased red tide blooms? hydrobiologia 352:141-147. holmes mj, gillespie nc, lewis rj, 1988. toxicity and morphology of ostreopsis cf. siamensis cultured from a ciguatera endemic region of queensland, australia, pp. 49-54. in: j.h. choat, d. barnes, m.a. borowitzka, j.c. coll, p.j. davies, p. flood, b.g. hatcher, d. hopley, p.a. hutchings, d. kinsey, g.r. orme, m. pichon, p.f. sale, p. sammarco, c.c. wallace, c. wilkinson, e. wolanski and o. bellwood (eds.), proc. 6th int. coral reef symp., townsville, australia. honsell g, de bortoli m, boscolo s, dell’aversano c, battocchi c, fontanive g, penna a, berti f, sosa s, yasumoto t, ciminiello p, poli m, tubaro a, 2011. harmful dinoflagellate ostreopsis cf. ovata fukuyo: detection of ovatoxins in field samples and cell immunolocalization using antipalytoxin antibodies. environ. sci. technol. 45:7051-7059. hoppenrath m, murray sa, chomérat n, horiguchi t, 2014. marine benthic dinoflagellates unveiling their worldwide biodiversity. schweizerbart science, stuttgart: 276 pp. illoul h, hernandez fr, vila m, adjas n, younes aa, bournissa m, koroghli a, marouf n, rabia s, ameur flk, 2012. the genus ostreopsis along the algerian coastal waters (sw mediterranean sea) associated with a human respiratory intoxication episode. cryptogam. algol. 33:209-216. ismael a, halim y, 2012. potentially harmful ostreopsis spp. in the coastal waters of alexandria-egypt. mediterr. mar. sci. 13:208-212. ispra, 2012. [ostreopsis cf. ovata lungo le coste italiane: monitoraggio 2011].[report in italian]. rapporti 173/2012. ispra. available from: http://www.isprambiente.gov.it/en/ publications/reports/ostreopsis-ovata-along-the-italian-coastmonitoring-2011?set_language=en jin q, dong s, 2003. comparative studies on the allelopathic effects of two different strains of ulva pertusa on heterosigma akashiwo and alexandrium tamarense. j. exp. mar. biol. ecol. 293:41-55. johansson n, granéli e, 1999a. cell density, chemical composition and toxicity of chrysochromulina polylepis (haptophyta) in relation to different n:p supply ratios. mar. biol. 135:209-217. johansson n, granéli e, 1999b. influence of different nutrient conditions on cell density, chemical composition and toxicity of prymnesium parvum (haptophyta) in semi-continuous cultures. j. exp. mar. biol. ecol. 239:243-258. kermarec f, dor f, armengaud a, charlet f, kantin r, sauzade d, de haro l, 2008. [les risques sanitaires liés à la présence d’ostreopsis ovata dans les eaux de baignade ou d’activités nautiques].[article in french]. environ risque sante 7:357-363. kim hs, yih w, kim jh, myung g, jeong hj, 2011. abundance of epiphytic dinoflagellates from coastal waters off jeju island, korea during autumn 2009. ocean sci. j. 46:205-209. kodama am, hokama y, yasumoto t, fukui m, manea sj, sutherland n, 1989. clinical and laboratory findings implicating palytoxin as cause of ciguatera poisoning due to decapterus macrosoma (mackerel). toxicon 27:1051-1053. laza-martinez a, orive e, miguel i, 2011. morphological and genetic characterization of benthic dinoflagellates of the genera coolia, ostreopsis and prorocentrum from the southeastern bay of biscay. eur. j. phycol. 46:45-65. leaw c-p, lim pt, ahmad a, usup g, 2001. genetic diversity of ostreopsis ovata (dinophyceae) from malaysia. mar. biotechnol. 3:246-255. lenoir s, ten-hage l, turquet j, quod jp, bernard c, hennion mc, 2004. first evidence of palytoxin analogues from an ostreopsis mascarenensis (dinophyceae) benthic bloom in southwestern indian ocean. j. phycol. 40:1042-1051. litaker rw, vandersea mw, kibler sr, madden vj, noga ej, tester pa, 2002. life cycle of the heterotrophic dinoflagellate pfiesteria piscicida (dinophyceae). j. phycol. 38:442-463. lobel ps, anderson dm, durand-clement m, 1988. assessment of ciguatera dinoflagellate populations sample variability and algal substrate selection. biol. bull. 175:94-101. mabrouk l, hamza a, brahim mb, bradai mn, 2011. temporal and depth distribution of microepiphytes on posidonia oceanica (l.) delile leaves in a meadow off tunisia. mar. ecol. 32:148-161. mabrouk l, hamza a, mahfoudi m, bradai mn, 2012. spatial and temporal variations of epiphytic ostreopsis siamensis on posidonia oceanica (l.) delile leaves in mahdia (tunisia). cah. biol. mar. 53:419-427. mangialajo l, bertolotto r, cattaneo-vietti r, chiantore m, grillo c, lemée r, melchiorre n, moretto p, povero p, ruggieri n, 2008. the toxic benthic dinoflagellate ostreopsis ovata: quantification of proliferation along the coastline of genoa, italy. mar. pollut. bull. 56:1209-1214. mangialajo l, ganzin n, accoroni s, asnaghi v, blanfuné a, cabrini m, cattaneo-vietti r, chavanon f, chiantore m, cohu s, costa e, fornasaro d, grossel h, marco-miralles f, masó m, reñé a, rossi am, sala mm, thibaut t, totti c, vila m, lemée r, 2011. trends in ostreopsis proliferation along the northern mediterranean coasts. toxicon 57:408-420. mercado ja, rivera-rentas al, gonzalez i, tosteson tr, molgo j, escalona de motta g, 1994. neuro-and myo-toxicity of extracts from the benthic dinoflagellate ostreopsis lenticularis is sensitive to µ-conotoxin. soc. neurosci. abstracts 20:718. no nco mm er cia l u se on ly ostreopsis in temperate areas meunier fa, mercado ja, molgó j, tosteson tr, escalona de motta g, 1997. selective depolarization of the muscle membrane in frog nerve-muscle preparations by a chromatographically purified extract of the dinoflagellate ostreopsis lenticularis. brit. j. pharmacol. 121:1224-1230. mizumoto h, tomaru y, takao y, shirai y, nagasaki k, 2008. diverse responses of the bivalve-killing dinoflagellate heterocapsa circularisquama to infection by a single-stranded rna virus. appl. environ. microbiol. 74:3105-3111. monti m, cecchin e, 2012. comparative growth of three strains of ostreopsis ovata at different light intensities with focus on inter-specific allelopathic interactions. cryptogam. algol. 33:113-119. monti m, minocci m, beran a, l. i, 2007. first record of ostreopsis cfr. ovata on macroalgae in the northern adriatic sea. mar. pollut. bull. 54:598-601. moore re, scheuer pj, 1971. palytoin: a new marine toxin from a coelenterate. science 172:495-498. moore re, bartolini g, 1981. structure of palytoxin. j. am. chem. soc. 103:2491-2494. morton sl, norris dr, bomber jw, 1992. effect of temperature, salinity and light-intensity on the growth and seasonality of toxic dinoflagellates associated with ciguatera. j. exp. mar. biol. ecol. 157:79-90. munday r, 2011. palytoxin toxicology: animal studies. toxicon 57:470-477. murray s, 2010. benthic dinoflagellates, p. 213-259. in: g. hallegraeff, c.j.s. bolch, d.r.a. hill, i. jameson, j.-m. leroi, a. mcminn, s. murray, m.f. de salas and k. saunders (eds.), phytoplankton flora of temperate australia. csiro publishing. nagasaki k, ando m, imai i, itakura s, ishida y, 1994. viruslike particles in heterosigma akashiwo (raphidophyceae): a possible red tide disintegration mechanism. mar. biol. 119:307-312. nagasaki k, yamaguchi m, imai i, 2000. algicidal activity of a killer bacterium against the harmful red tide dinoflagellate heterocapsa circularisquama isolated from ago bay, japan. nippon suisan gakkaishi 66:666-673. nakajima i, oshima y, yasumoto t, 1981. toxicity of benthic dinoflagellates in okinawa. bull. japan soc. sci. fish. 47:1029-1033. norris dr, bomber jw, balech e, 1985. benthic dinoflagellates associated with ciguatera from florida keys. i. ostreopsis heptagona sp. nov, pp. 39–44. in: d.m. anderson, a.w. white and d.g. baden (eds.), toxic dinoflagellates. elsevier. okolodkov yb, campos-bautista g, gárate-lizárraga i, gonzález-gonzález jag, hoppenrath m, arenas v, 2007. seasonal changes of benthic and epiphytic dinoflagellates in the veracruz reef zone, gulf of mexico. aquat. microb. ecol. 47:223-237. onuma y, satake m, ukena t, roux j, chanteau s, rasolofonirina n, ratsimaloto m, naoki h, yasumoto t, 1999. identification of putative palytoxin as the cause of clupeotoxism. toxicon 37:55-65. pagliara p, caroppo c, 2012. toxicity assessment of amphidinium carterae, coolia cfr. monotis and ostreopsis cfr. ovata (dinophyta) isolated from the northern ionian sea (mediterranean sea). toxicon 60:1203-1214. parsons ml, preskitt lb, 2007. a survey of epiphytic dinoflagellates from the coastal waters of the island of hawai’i. harmful algae 6:658-669. pearce i, marshall ja, hallegraeff g, 2001. toxic epiphytic dinoflagellates from east coast tasmania, australia, pp. 54– 57. in: g. hallegraeff, s.i. blackburn, c.j. bolch and r.j. lewis (eds.), harmful algal blooms. intergovernmental oceanographic commission of unesco. penna a, vila m, fraga s, giacobbe mg, andreoni f, riobó p, vernesi c, 2005. characterization of ostreopsis and coolia (dinophyceae) isolates in the western mediterranean sea based on morphology, toxicity and internal transcribed spacer 5.8s rdna sequences. j. phycol. 41:212-225. penna a, fraga s, battocchi c, casabianca s, giacobbe mg, riobó p, vernesi c, 2010. a phylogeographical study of the toxic benthic dinoflagellate genus ostreopsis schmidt. j. biogeogr. 37:830-841. penna a, fraga s, battocchi c, casabianca s, perini f, samuela c, casabianca a, riobo p, giacobbe m, totti c, accoroni s, vila m, reñé a, scardi m, aligizaki k, nguyen-ngoc l, vernesi c, 2012. genetic diversity of the genus ostreopsis schmidt: phylogeographical considerations and molecular methodology applications for field detection in the mediterranean sea. cryptogam. algol. 33:153-163. penna a, battocchi c, capellacci s, fraga s, aligizaki k, lemée r, vernesi c, 2014. mitochondrial, but not rdna, genes fail to discriminate dinoflagellate species in the genus ostreopsis. harmful algae 40:40-50. perini f, casabianca a, battocchi c, accoroni s, totti c, penna a, 2011. new approach using the real-time pcr method for estimation of the toxic marine dinoflagellate ostreopsis cf. ovata in marine environment. plos one 6:e17699. pezzolesi l, guerrini f, ciminiello p, dell’aversano c, dello iacovo e, fattorusso e, forino m, tartaglione l, pistocchi r, 2012. influence of temperature and salinity on ostreopsis cf. ovata growth and evaluation of toxin content through hr lc-ms and biological assays. water res. 46:82-92. pezzolesi l, pistocchi r, fratangeli f, dell’aversano c, dello iacovo e, tartaglione l, 2014. growth dynamics in relation to the production of the main cellular components in the toxic dinoflagellate ostreopsis cf. ovata. harmful algae 36:1-10. pfannkuchen m, godrijan j, maric pfannkuchen d, ivesa l, kruzic p, ciminiello p, dell’aversano c, dello iacovo e, fattorusso e, forino m, tartaglione l, godrijan m, 2012. toxinproducing ostreopsis cf. ovata are likely to bloom undetected along coastal areas. environ. sci. technol. 46:5574-5582. pfiester la, 1989. dinoflagellate sexuality. int. rev. cytol. 114:249-272. pfiester la, anderson dm, 1987. dinoflagellate reproduction, p. 611-648. in: f.j.r. taylor (ed.), the biology of dinoflagellates. blackwell science. pichierri s, pezzolesi l, accoroni s, guerrini f, pistocchi r, totti c, 2015. growth inhibition of ostreopsis cf. ovata culture by toxic algal-derived polyunsaturated aldehydes. proc. cx nat. congr. società botanica italiana, pavia, italy. pinna a, pezzolesi l, pistocchi r, vanucci s, ciavatta s, polimene l, 2015. modelling the stoichiometric regulation of c-rich toxins in marine dinoflagellates. plos one 10, e0139046. pistocchi r, pezzolesi l, guerrini f, vanucci s, dell’aversano c, fattorusso e, 2011. a review on the effects of environ13 no nco mm er cia l u se on ly 14 s. accoroni and c. totti mental conditions on growth and toxin production of ostreopsis ovata. toxicon 57:421-428. privitera d, giussani v, isola g, faimali m, piazza v, garaventa f, asnaghi v, cantamessa e, cattaneo-vietti r, chiantore m, 2012. toxic effects of ostreopsis ovata on larvae and juveniles of paracentrotus lividus. harmful algae 18:16-23. quod jp, 1994. ostreopsis mascarenensis sp. nov. (dinophyceae), dinoflagellé toxique associé a la ciguatera dans l‘océan indien. cryptogam. algol. 15:243-251. randall je, 2005. review of clupeotoxism, an often fatal illness from the consumption of clupeoid fishes. pac. sci. 59:73-77. raniello r, lorenti m, brunet c, buia mc, 2004. photosynthetic plasticity of an invasive variety of caulerpa racemosa in a coastal mediterranean area: light harvesting capacity and seasonal acclimation. mar. ecol.-prog. ser. 271:113-120. rhodes l, adamson j, suzuki t, briggs l, garthwaite i, 2000. toxic marine epiphytic dinoflagellates, ostreopsis siamensis and coolia monotis (dinophyceae), in new zealand. new zeal. j. mar. fresh. 34:371-383. rhodes l, towers n, briggs l, munday r, adamson j, 2002. uptake of palytoxin-like compounds by shellfish fed ostreopsis siamensis (dinophyceae). n. z. j. mar. freshw. res. 36:631-636. rhodes ll, smith kf, munday r, selwood ai, mcnabb ps, holland pt, bottein m-y, 2010. toxic dinoflagellates (dinophyceae) from rarotonga, cook islands. toxicon 56:751-758. richlen ml, lobel ps, 2011. effects of depth, habitat, and water motion on the abundance and distribution of ciguatera dinoflagellates at johnston atoll, pacific ocean mar. ecol.prog. ser. 421:51-66. riobó p, paz b, franco jm, 2006. analysis of palytoxin-like in ostreopsis cultures by liquid chromatography with precolumn derivatization and fluorescence detection. anal. chim. acta 566:217-223. riobó p, paz b, franco jm, vázquez ja, murado ma, 2008. proposal for a simple and sensitive haemolytic assay for palytoxin toxicological dynamics, kinetics, ouabain inhibition and thermal stability. harmful algae 7:415-429. rossi r, castellano v, scalco e, serpe l, zingone a, soprano v, 2010. new palytoxin-like molecules in mediterranean ostreopsis cf. ovata (dinoflagellates) and in palythoa tuberculosa detected by liquid chromatography-electrospray ionization time-of-flight mass spectrometry. toxicon 56:1381-1387. rossini gp, bigiani a, 2011. palytoxin action on the na+,k+atpase and the disruption of ion equilibria in biological systems. toxicon 57:429-439. sagara t, 2008. profiles of palytoxin-like compounds from the dinoflagellate ostreopsis sp. isolated from the areas where poisonous fishes were collected. nippon suisan gakkaishi 74:913-914. scalco e, brunet c, marino f, rossi r, soprano v, zingone a, montresor m, 2012. growth and toxicity responses of mediterranean ostreopsis cf. ovata to seasonal irradiance and temperature conditions. harmful algae 17:25-34. schmidt j, 1901. flora of koh chang. contributions to the knowledge of the vegetation in the gulf of siam. peridiniales. botanische tidsskrift 24:212-221. selina ms, orlova ty, 2010. first occurrence of the genus ostreopsis (dinophyceae) in the sea of japan. bot. mar. 53:243-249. selina ms, levchenko ev, 2011. species composition and morphology of dinoflagellates (dinophyta) of epiphytic assemblages of peter the great bay in the sea of japan. russ. j. mar. biol. 37:23-32. selina ms, morozova tv, vyshkvartsev di, orlova ty, 2014. seasonal dynamics and spatial distribution of epiphytic dinoflagellates in peter the great bay (sea of japan) with special emphasis on ostreopsis species. harmful algae 32:1-10. shears nt, ross pm, 2009. blooms of benthic dinoflagellates of the genus ostreopsis; an increasing and ecologically important phenomenon on temperate reefs in new zealand and worldwide. harmful algae 8:916-925. shears nt, ross pm, 2010. toxic cascades: multiple anthropogenic stressors have complex and unanticipated interactive effects on temperate reefs. ecol. lett. 13:1149-1159. simonini r, orlandi m, abbate m, 2011. is the toxic dinoflagellate ostreopsis cf. ovata harmful to mediterranean benthic invertebrates? evidences from ecotoxicological tests with the polychaete dinophilus gyrociliatus. mar. environ. res. 72:230-233. sosa s, del favero g, de bortoli m, vita f, soranzo mr, beltramo d, ardizzone m, tubaro a, 2009. palytoxin toxicity after acute oral administration in mice. toxicol. lett. 191:253-259. spatharis s, dolapsakis np, economou-amilli a, tsirtsis g, danielidis db, 2009. dynamics of potentially harmful microalgae in a confined mediterranean gulf-assessing the risk of bloom formation. harmful algae 8:736-743. taniyama s, arakawa o, terada m, nishio s, takatani t, mahmud y, noguchi t, 2003. ostreopsis sp., a possible origin of palytoxin (ptx) in parrotfish scarus ovifrons. toxicon 42:29-33. tarutani k, nagasaki k, itakura s, yamaguchi m, 2001. isolation of a virus infecting the novel shellfish-killing dinoflagellate heterocapsa circularisquama. aquat. microb. ecol. 23:103-111. tawong w, yoshimatsu t, yamaguchi h, adachi m, 2015. effects of temperature, salinity and their interaction on growth of benthic dinoflagellates ostreopsis spp. from thailand. harmful algae 44:37-45. taylor fjr, 1979. the description of the benthic dinoflagellate associated with maitotoxin and ciguatoxin, including observations on hawaiian material, p. 71-77. in: d.l. taylor and h.h. seliger (eds.), toxic dinoflagellate blooms. elsevier scientific. tichadou l, glaizal m, armengaud a, grossel h, lemée r, kantin r, lasalle jl, drouet g, rambaud l, malfait p, de haro l, 2010. health impact of unicellular algae of the ostreopsis genus blooms in the mediterranean sea: experience of the french mediterranean coast surveillance network from 2006 to 2009. clinical toxicology 48:839-844. tognetto l, bellato s, moro i, andreoli c, 1995. occurrence of ostreopsis ovata (dinophyceae) in the tyrrhenian sea during summer 1994. bot. mar. 38:291-295. tosteson tr, 1995. the diversity and origins of toxins in ciguatera fish poisoning. p. r. health sci. j. 14:117-129. totti c, accoroni s, cerino f, cucchiari e, romagnoli t, 2010. ostreopsis ovata bloom along the conero riviera (northern adriatic sea): relationships with environmental conditions and substrata. harmful algae 9:233-239. no nco mm er cia l u se on ly ostreopsis in temperate areas tubaro a, del favero g, beltramo d, ardizzone m, forino m, de bortoli m, pelin m, poli m, bignami g, ciminiello p, sosa s, 2011. acute oral toxicity in mice of a new palytoxin analog: 42-hydroxy-palytoxin. toxicon 57:755-763. tubaro a, del favero g, pelin m, bignami g, poli m, 2014. palytoxin and analogues: biological effects and detection, p. 741-772. in: l.m. botana (ed.), seafood and freshwater toxins: pharmacology, physiology, and detection, 3rd edition. crc press. turki s, 2005. distribution of toxic dinoflagellates along the leaves of seagrass posidonia oceanica and cymodocea nodosa from the gulf of tunis. cah. biol. mar. 46:29-34. turki s, harzallah a, sammari c, 2006. occurrence of harmful dinoflagellates in two different tunisian ecosystems: the lake of bizerte and the gulf of gabes. cah. biol. mar. 47:253-259. uchida h, taira y, yasumoto t, 2013. structural elucidation of palytoxin analogs produced by the dinoflagellate ostreopsis ovata ik2 strain by complementary use of positive and negative ion liquid chromatography/quadrupole time-of-flight mass spectrometry. rapid commun. mass spectrom. 27:1999-2008. ukena t, satake m, usami m, oshima y, naoki h, fujita t, kan y, yasumoto t, 2001. structure elucidation of ostreocin d, a palytoxin analog isolated from the dinoflagellate ostreopsis siamensis. biosci. biotechnol. biochem. 65:2585-2588. ungaro n, marano g, pastorelli am, marzano mc, pompei m, 2005. presenza di ostreopsidiaceae nel basso adriatico, rapporti istisan 05/29. istituto superiore di sanità, roma. usami m, satake m, ishida s, inoue a, kan y, yasumoto t, 1995. palytoxin analogs from the dinoflagellate ostreopsis siamensis. j. am. chem. soc. 117:5389-5390. vanucci s, guerrini f, pezzolesi l, dell’aversano c, ciminiello p, pistocchi r, 2012a. cell growth and toxins’ content of ostreopsis cf. ovata in presence and absence of associated bacteria. cryptogam. algol. 33:105-112. vanucci s, pezzolesi l, pistocchi r, ciminiello p, dell’aversano c, dello iacovo e, fattorusso e, tartaglione l, guerrini f, 2012b. nitrogen and phosphorus limitation effects on cell growth, biovolume, and toxin production in ostreopsis cf. ovata. harmful algae 15:78-90. vidyarathna nk, granéli e, 2013. physiological responses of ostreopsis ovata to changes in n and p availability and temperature increase. harmful algae 21-22:54-63. vila m, garcés e, masó m, 2001. potentially toxic epiphytic dinoflagellate assemblages on macroalgae in the nw mediterranean. aquat. microb. ecol. 26(1):51-60. wall d, 1975. taxonomy and cysts of red-tide dinoflagellates, p. 249-255. in: v.r. locicero (ed.), first international conference on toxic dinoflagellate blooms. massachusetts science and technologyfoundation, wakefield. wattenberg ev, 2011. modulation of protein kinase signaling cascades by palytoxin. toxicon 57:440-448. yasumoto t, 1998. fish poisoning due to toxins of microalgal origins in the pacific. toxicon 36:1515-1518. yasumoto t, yasumura d, ohizumi y, takahashi m, alcala ac, alcala lc, 1986. palytoxin in two species of xanthid crab from the philippines. agric. biol. chem. 50:163-167. yasumoto t, seino n, murakami y, murata m, 1987. toxins produced by benthic dinoflagellates. biol. bull. 172:128-131. yamaguchi h, tomori y, tanimoto y, oku o, adachi m, 2014. evaluation of the effects of light intensity on growth of the benthic dinoflagellate ostreopsis sp. 1 using a newly developed photoirradiation-culture system and a novel regression analytical method. harmful algae 39:48-54. zhang q, hu g, 2011. effect of nitrogen to phosphorus ratios on cell proliferation in marine micro algae. chin. j. oceanol. limnol. 29:739-745. 15 no nco mm er cia l u se on ly layout 1 introduction the use of chemical cues by animals is widespread in all environments, including freshwater habitats, where changes in behaviour after the detection of waterborne compounds have been extensively documented for many species and in many behavioural contexts (wyatt, 2014). like other decapod crustaceans, crayfish rely on waterborne chemical signals to produce appropriate behavioural responses, such as spatial orientation, social communication, predator avoidance, identification of suitable habitats, location of food resources and sex recognition (breithaupt and eger, 2002; grasso and basil, 2002; bergman and moore, 2005; aquiloni et al., 2009; berry and breithaupt, 2010; breithaupt, 2011; schmidt and mellon, 2011; solari et al., 2015, 2017, 2018). discrimination of different stimuli is mediated by peripheral chemoreceptors grouped within sensory hairs called sensilla, which are typically located on the cuticles of cephalothoracic appendages like antennae, maxillipeds (mouthparts) and especially antennules and pereiopods (major claws and walking legs) (schmidt and mellon, 2011). pheromone-induced sex recognition, in particular, has been widely documented in different crayfish species and supports the involvement of pheromones in some aspects of crayfish sexual behaviour. for instance, pioneering studies have reported chemoreception and pheromones for sex discrimination in the red swamp crayfish procambarus clarkii (girard, 1852) having intact and functional antennules (ameyaw-akumfi and hazlett, 1975, dunham and oh, 1992). chemical detection of sex was reported for the genus orconectes as well (hazlett, 1985; simon and moore, 2007), and the presence of a sex pheromone released by mature females to stimulate courtship and mating behaviour in males was also described in the signal crayfish pacifastacus leniusculus (stebbing et al., 2003). additionally, by using urine blocking protocols and visualization techniques with fluorescein dye, urine released anteriorly through a pair of nephropores into water currents generated by the animals themselves was shown to be the source of the pheromone signals (simon and moore, 2007; berry and breithaupt, 2010). in decapods, as in other crustaceans, the nephropores can be opened and closed by a sphincter muscle (bushmann and atema, 1996; subramoniam, 2016) controlling modality and timing of pheromone release. chemoreceptors housed in the antennular flagella were reported to be, in most cases, responsible for pheromone detection (tierney et al., 1984); the major chelae were also suggested as possible, alternative sites of sex pheromone discrimination (belanger and moore 2006). however, a number of studies on the use of pheromones in crayfish mating behaviour produced contradictory results, suggesting that crayfish do not necessarily communicate their sex chemically (itagaki and thorp, 1981; thorp, 1984). other studies on p. clarkii, based on ablation experiments in both males and females, showed that the presence of antennules is not necessary for mating behaviour (corotto et al.,1999). recent evidence suggests that females of p. clarkii make use of combined visual and chemical information to more easily identify larger males and to discriminate between individuals of the opposite sex, and also tactile sensory input during physical interactions seems to provide crucial information to the female for assessment of male quality (aquiloni and gherardi, 2008; aquiloni et al., 2009). article the success in the short-distance communication for mating does not depend on chemical signals in the crustacean decapod procambarus clarkii (girard, 1852) stefania peddio,1 giorgia sollai,1 cinzia podda,2 giacomo frau,2 francesco palmas,2 andrea sabatini,2 roberto crnjar,1 paolo solari1* 1department of biomedical sciences, section of physiology, university of cagliari, university campus, s.p. 8, 09042 monserrato (ca), italy; 2department of life and environmental sciences, university of cagliari, via fiorelli 1, 09126 cagliari, italy abstract pheromone-driven sex recognition has been widely documented in crayfish and a great deal of evidence supports the involvement of pheromones in their mating behaviour. this study investigates whether sexual interaction and mating success in the red swamp crayfish procambarus clarkii are dependent on short-distance chemical communication between sexes, mediated by urine-borne pheromones. we compared the mating behaviour of intact animals that could release urine to chemically communicate in a natural way with that of urine-blocked animal pairs, for which chemical communication was precluded. our results show that urine-borne pheromones are not critical for the reproductive success of p. clarkii, at least over the short-range distance (<1 m) considered in this study, during which the animals were confined in a restricted tank, facing one each other, and thus able to promptly mate. under these experimental conditions, a lack of urine release neither precluded the occurrence, nor affected the duration of the different phases of mating behaviour. we conclude that short-distance chemical communication in p. clarkii is not a prerequisite for mating, and suggest that it could be affected by alternative sensory modalities, likely vision and/or acoustic signalling. no nco mm er cia l u se on ly s. peddio et al.68 based on these considerations and given the lack of information in the literature about the spatial range of pheromone action, the aim of the present study was to ascertain whether mating success in the red swamp crayfish is strictly dependent on short-distance communication between sexes, as mediated by urineborne pheromones; that is, when the animals are in close proximity in a confined environment and therefore ready to start sexual interactions. specifically, we measured the duration of the different phasesof reproductive behaviour in intact animals, which could eventually release urine in order to chemically communicate in a natural way (control group), and compared the results with the duration of the corresponding phases in urine-blocked animals, for which chemical communication was precluded. methods animal collection and rearing conditions adult red swamp crayfish p. clarkii of both sexes, 3540 mm in carapace length, were collected using a backpack electrofishing unit (5.2-2.8 a, 230-400 v, 1300 w) in molentargius-saline regional natural park (southern sardinia, italy) during the spring season of 2018-2019. the crayfish were kept in plexiglas® tanks (100 cm long, 50 cm wide, 20 cm deep) containing 60 l of aerated and bio-conditioned (aquasafe, tetra, melle, germany) tap water (hereafter referred to as tap water), at 22-23°c, 16 h light/8 h dark photoperiodic regime, and fed with lettuce, squid or a highly appetitive commercial pellet food (shrimps natural, sera, heinsberg, germany) three times a week. uneaten food was always removed within 1 h of delivery. individuals were kept separate to avoid any reciprocal exposure of males and females and to prevent attacks or cannibalism. urine blocking technique to examine the role that the urine-borne compounds may play in sexual interactions occurring during reproduction, with special regard to the pheromone(s), treated crayfish nephropores were closed, thus blocking urine release, in both sexes 3 h prior to experimentation, according to the procedure adopted for pacifastacus leniusculus (berry and breithaupt, 2010). practically, crayfish were removed from their tanks and immobilised, and nephropores were blocked by attaching a small piece of silicon tubing (5 mm length x 1.9 mm diameter) to the shell surface surrounding each nephropore using cyanoacrylate glue (loctite, super attak power flex). to be sure on the tubing functionality, an additional cyanoacrylate layer was applied around the tube and allowed to dry. the open end of the tubing was sealed with a plasticine plug and reinforced with cyanocrylate glue. dye tests were also performed in order to ensure that the nephropore block was sufficient to prevent any urine release during crayfish interactions. all plugs were removed at the end of the experiments and crayfish were transferred to the holding tank. each individual was used in only one mating experiment. animals were not fed for a 24 h period preceding the experiments. the sexual maturity of the crayfish was assessed by the presence of well-developed glair glands in females and whitened gonopods in males (mclay and van den brink, 2016). experimental protocol individuals of each sex were introduced to either side of a plexiglas® tank (40 cm long, 30 cm wide, 15 cm deep) containing 10 l of tap water (22-23 °c), separated by a central acrylic divider, and left to acclimatise for 15 min. the divider was then lifted and animals were allowed to interact for a maximum of 4 h. following each experiment, the tanks and dividers were washed thoroughly. trials were video-recorded for later analysis using a samsung smx-f34 (samsung, seoul, korea) colour digital camera mounted above the test tank. video recordings were analysed by an independent observer blind to the experimental treatment. the courtship behaviour of both the open-nephropore crayfish (n = 9 pairs) and the blocked-nephropore animals (n = 8 pairs) was recorded, analysed and categorised into the following five phases (hereafter abbreviated as ph1 ph5; fig. 1; stebbing et al., 2003): ph1. orientation: the male oriented and started moving towards the female, which tended to remain stationary until the male approached; ph2. contact: the animals met with chelae raised and the female became submissive. mating contact was clearly distinguishable from non-mating contact, because the female did not swim away when grasped by the male; ph3. pre-copulation mounting: the male grasped the female by her chelipeds, antennae or rostrum in order to climb onto her back, turn her over and to move along her body until their ventral surfaces were opposed; ph4. spermatophore deposition: the male deposited spermatophores onto the female’s ventral surface; ph5. dismounting: after deposition, the females struggled and turned over, throwing the males off and moving away. to assess the effect of the urine block, and therefore of the lack of pheromone communication, on the sexual interactions and mating success, the duration of each phase was measured and compared between the two experimental groups. no nco mm er cia l u se on ly mating behaviour in the red swamp crayfish 69 all data (mean time ± se) were subjected to one-way analysis of variance (anova) with a 95% confidence level (p<0.05). data were first checked for normality and homogeneity of variance using the kolmogorov-smirnov and the levene tests, respectively, which allowed us to use parametric tests. statistical analyses were carried out using statistica for windows (ver. 7.0; statsoft inc., tulsa, ok, usa). results in terms of the duration of each of the five phases of crayfish mating behaviour (fig. 2), we found that blocking urine release did not affect animal orientation (ph1) (f[1.15] = 0.0916; p=0.77), which is the initial, critical phase of the mating process, in which the crayfish begin to approach each other. in p. clarkii, this proved to be a relatively fast phase and lasted only a few minutes for both the control (2.99 ± 0.69 min) and the urine-blocked animals (3.28 ± 0.70 min). the following phase, the contact between animals (ph2), was longer than the orientation phase, and was characterised by the fact that, before becoming submissive, the female exhibited an initial level of aggressiveness towards the male. despite the great variability in the response exhibited by both of the experimental groups of animals, the duration of ph2 was not affected by a lack of urine release (f[1.15] = 0.6841; p=0.42). the third phase considered was the pre-copulation mounting phase (ph3), starting with the immobilisation fig. 1. representative pictures showing the different phases of the p. clarkii mating behaviour considered in this study, orientation (a), contact (b), pre-copulation mounting (c), spermatophore deposition (d) and dismounting (e). no nco mm er cia l u se on ly s. peddio et al.70 of the female and culminating in the opposition of the animals’ ventral surfaces. overall, ph3 was relatively fast and lasted only 0.63 ± 0.10 and 0.56 ± 0.05 min in the control animals and in urine-blocked ones, respectively. like previous phases, ph3 did not depend on the release of urine and therefore on the ability of the animals to communicate chemically with each other (f[1.15] = 0.2997; p=0.59). the fourth phase (ph4), which is most critical for the success of reproductive behaviour, consisted of the gonopod-guided deposition of the male spermatophores within the annulus ventralis of the female. ph4 proved to be long indeed, lasting 35.77 ± 5.20 min in the control pairs and 39.19 ± 7.93 min in the blocked-nephropore ones, but once again, blocking urine release did not produce any significant effect (f[1.15] = 0.1361; p=0.72). the final dismounting phase (ph5) was also unaffected by a lack of chemical communication between sexes (f[1.15] = 0.4879; p=0.49); females from the urineblocked group were able to struggle and turn over, throwing the males off, in a time that was comparable with that of control females. discussion and conclusions the present study shows that in the red swamp crayfish p. clarkii, urine-borne sex pheromones are not critical for interactions between sexes aimed at the reproductive success, at least on the small spatial scale considered; the animals were confined in a restricted environment, facing each other and ready to start mating behaviour. in fact, under these experimental conditions, the lack of urine release neither precluded the occurrence, nor affected the duration of the different phases of the crayfish mating, as the behavioural patterns of urineblocked animals were comparable to those of control crayfish, which could release their urine and therefore chemically communicate in a natural way. if the ability of these animals to identify the sex of a conspecific and to reciprocally interact for mating purposes was impaired by a lack of urine-borne chemicals, one might expect it to take longer for an animal to accomplish these tasks, therefore delaying mating or, eventually, preventing it. it is likely that the initial approach, including orientation and the initial contact between animals, are the most critical moments of sex recognition and for obtaining consent to mate, so it is in these phases that the ability to chemically communicate might play a decisive role. the duration of the contact phase, in particular, seemed to be inversely proportional to the degree of affinity between the two sexes, in such a way that the higher was the affinity between the partners, the shorter was the time to complete this phase. in the case of p. clarkii, neither of these phases were significantly delayed by blocking urine, suggesting that short-distance urine-borne pheromones were not necessary for mating in these animals. in this respect, p. clarkiidiffers fig. 2. time spent (min) during the different phases of the mating behaviour, orientation (ph1), contact (ph2), pre-copulation mounting (ph3), spermatophore deposition (ph4) and dismounting (ph5) in the intact (control) and in the urine-blocked crayfish pairs. values are means ± sem (vertical bars) from 9 intact and 8 urine-blocked animal pairs. no nco mm er cia l u se on ly mating behaviour in the red swamp crayfish 71 from other freshwater crayfish like p. leniusculus. in fact, in the case of p. leniusculus, unambiguous responses such as “mating” of males with an air-stone releasing conditioned water from mature females clearly pointed to the involvement of a pheromone that induces courtship and mating behaviour in males (stebbing et al., 2003). furthermore, in p. leniusculus, blocking female urination prevented any male courtship behaviour, while artificial introduction of female urine was found to re-establish male mating attempts (berry and breithaupt, 2010). in the case of p. clarkii, previous experimental observations are contradictory; a number of studies documented the use of chemoreception and pheromones for sex discrimination, but only if the antennules were intact (ameyaw-akumfi and hazlett, 1975, dunham and oh, 1992). on the other hand, it must be noted that p. clarkii neither communicate their sex chemically, nor does antennule loss measurably impair sex discrimination, at least when the animals are in close proximity (itagaki and thorp, 1981; thorp, 1984; corotto et al., 1999). disregarding the different experimental designs that gave rise to these discrepancies, our findings support the hypothesis that sex discrimination and mating behaviour in p. clarkii do not necessarily involve urine-borne chemicals. removing antennules might not necessarily lead to the same conclusions, as crayfish can use non-antennule chemoreception for sex discrimination, as suggested by corotto et al. (1999) in p. clarkii and belanger and moore (2006) in o. rusticus. certainly, we cannot exclude the possibility that sex pheromones and chemical communication in p. clarkii are involved in some aspects of the sexual behaviour. this could especially hold true over the long distances, over which olfactory cues may support searching behaviour better than any other sensory modality. this is a wellknown, pheromone-guided, stereotyped behaviour for mate search, typical of many animals, including crustaceans and insects, in which it is often referred to as “calling behaviour” (kaissling, 2014). however, when the crayfish are already in close proximity and facing one another, as in our experimental conditions, olfaction could be of less importance for sex discrimination and even redundant or easily overtaken in importance by alternative, faster methods of communication, such as visual, acoustic, or multimodal sensory inputs. in fact, it was suggested that even if they are most active at night (penn, 1943), females of p. clarkii are able to combine visual and chemical information to readily identify large males and to discriminate among individuals of opposite sex; tactile input during physical interactions also seems to provide crucial information to the female for assessment of male quality (aquiloni and gherardi, 2008; aquiloni et al., 2009). similarly, it was reported that p. clarkii can communicate underwater by producing and detecting sounds (buscaino et al., 2012); it is likely that this ability can also be used for sex discrimination during the initial phases of mating when the animals are already in close proximity, with no need for additional chemical communication. further investigation is needed to determine if alternative sensory modalities, likely vision and/or acoustics, are involved in these processes. our results also show that the three other phases of mating studied, pre-copulation mounting, spermatophore deposition and dismounting, occurred with no delay in urine-blocked animals compared to intact animals; that is, these phases occurred regardless of the ability of the animals to exchange chemical cues. this also suggests that once mating behaviour starts, it may proceed and can be completed via a stereotyped procedure that does not require further signals. on the whole, the five-stage mating behaviour of p. clarkii described here was substantially similar, from a qualitative point of view, to that exhibited by other crayfish species such as p. leniusculus, pacifastacus trowbridgii and austropotamobius pallipes, although the number of phases considered, three to seven, was different in the different species (mason, 1970; ingle and thomas, 1974; villanelli and gherardi, 1998; stebbing et al., 2003). in particular, the main phases of contact, turning, mounting, and spermatophore deposition were observed in all other crayfish species studied. conversely, the time that p. clarkii took to complete the mating process, about 60-80 min, was much longer than that observed in other crayfish, 5 min for a. pallipes (villanelli and gherardi, 1998), 8-20 min for p. trowbridgii (mason, 1970) and 2836 min in p. leniusculus (stebbing et al., 2003), as in p. clarkii the contact and the spermatophore deposition phases took a very long time. in conclusion, short-range sex discrimination and the occurrence of mating behaviour in the crayfish p. clarkii do not require chemical communication between sexes and, when the animals are in close proximity and ready to interact, they can begin and complete all phases of mating even when neither sex is able to release urineborne pheromones. at present, we cannot ascertain if this may represent an advantageous reproductive strategy for this crayfish, but we anticipate that it could represent a way to facilitate and optimise mating for a high yield. therefore, together with the considerable ecological plasticity, resistance, aggressiveness and bioturbation ability, this reproductive feature may increase the high invasive potential of p. clarkii for freshwater habitats worldwide (gherardi, 2006; salvadori et al., 2014, souty-grosset et al., 2016; palmas et al., 2019). for these reasons, this species is now listed among the “100 of the worst” invasive species by the “delivering alien invasive species inventories for europe” project (daisie, 2009), thus requiring severe population control strategies (manfrin et al., 2019). no nco mm er cia l u se on ly s. peddio et al.72 acknowledgments this research was partially supported by the “fondazione con il sud”, italy (grant no. 2015-0065). we thank marco melis, department of biomedical sciences, university of cagliari, for his helpful cooperation and technical support in the experimental part of the study. all applicable international, national, and/or institutional guidelines for the care and use of animals were followed. corresponding author: solari@unica.it key words: crayfish; invasive alien species; chemoreception; pheromones; olfaction; sex recognition. conflict of interest: the authors declare no competing interests. received: 11 october 2019. accepted: 3 december 2019. this work is licensed under a creative commons attribution noncommercial 4.0 license (cc by-nc 4.0). ©copyright: the author(s), 2019 licensee pagepress, italy advances in oceanography and limnology, 2019; 10:8617 doi: 10.4081/aiol.2019.8617 references ameyaw-akumfi c, hazlett ba, 1975. sex recognition in the crayfish procambarus clarkii. science 190:1225-1226. aquiloni l, gherardi f, 2008. assessing mate size in the red swamp crayfish procambarus clarkii: effects of visual versus chemical stimuli. freshwater biol. 53:461-469. aquiloni l, massolo a, gherardi f, 2009. sex identification in female crayfish is bimodal. naturwissenschaften 96:103110. belanger rm, moore pa, 2006. the use of the major chelae by reproductive male crayfish (orconectes rusticus) for discrimination of female odours. behaviour 143:713-731. bergman da, moore pa, 2005. the role of chemical signals in the social behavior of crayfish. chem. senses 30:305-306. berry fc, breithaupt t, 2010. to signal or not to signal? chemical communication by urine-borne signals mirrors sexual conflict in crayfish. bmc biol. 8:25. breithaupt t, eger p, 2002. urine makes the difference: chemical communication in fighting crayfish made visible. j. exp. biol. 205:1221-1231. breithaupt t, 2011. chemical communication in crayfish, p. 257-276. in: t. breithaupt and m. thiel (eds.), chemical communication in crustaceans. springer, new york. buscaino g, filiciotto f, buffa g, di stefano v, maccarrone v, buscaino c, mazzola s, alonge g, d’angelo s, maccarrone v, 2012. the underwater acoustic activities of the red swamp crayfish procambarus clarkii. j. acoust. soc. am. 132:1792-1798. bushmann p, atema j, 1996. nephropore rosette glands of the lobster homarus americanus: possible sources of urine pheromones. j. crust. biol. 16:221-231. corotto fs, bonenberger dm, bounkeo jm, dukas cc, 1999. antennule ablation, sex discrimination and mating behavior in the crayfish procambarus clarkii. j. crust. biol. 19:708712. daisie, 2009. handbook of alien species in europe. invading nature: springer series in invasion ecology, 3. springer, dordrecht: 399 pp. dunham dw, oh jw, 1992. chemical sex-discrimination in the crayfish procambarus clarkii: role of antennules. j. chem. ecol. 18:2363-2372. gherardi f, 2006. crayfish invading europe: the case study of procambarus clarkii. mar. freshw. behav. physiol. 39:175-191. grasso fw, basil ja, 2002. how lobsters, crayfishes, and crabs locate sources of odor: current perspectives and future directions. curr. opin. neurobiol. 12:721-727. hazlett ba, 1985. chemical detection of sex and condition in the crayfish orconectes virilis. j. chem. ecol. 11:181-189. ingle rw, thomas w, 1974. mating and spawning of the crayfish austropotamobius pallipes (crustacea: astacidae). j. zool. 173:525-538. itagaki h, thorp jh, 1981. laboratory experiments to determine if crayfish can communicate chemically in a flow-through system. j. chem. ecol. 7:115-126. kaissling ke, 2014. pheromone reception in insects (the example of silk moths), p. 99-146. in: c. mucignat-caretta (ed.), neurobiology of chemical communication. crc press, boca raton, fl, usa. manfrin c, souty-grosset c, anastácio pm, reynolds j, giulianini pg, 2019. detection and control of invasive freshwater crayfish: from traditional to innovative methods. diversity 11:5. mason jc, 1970. copulatory behaviour of the crayfish, pacifastacus trowbridgii (stimpson). can. j. zool. 48:969976. mclay cl, van den brink am, 2016. crayfish growth and reproduction, p. 62-116. in: m. longshaw and p. stebbing (eds), biology and ecology of crayfish. crc press, boca raton, fl, usa. palmas f, podda c, frau g, cau al, moccia d, peddio s, solari p, pusceddu a, sabatini a, 2019. invasive crayfish (procambarus clarkii, girard, 1852) in a managed brackish wetland (sardinia, italy): controlling factors and effects on sedimentary organic matter. estuar. coast. shelf sci. 231:106459. penn gh jr. 1943. a study of the life history of the louisiana red crayfish (girard). ecology 24:1-18. salvadori s, coluccia e, deidda f, cau a, cannas r, lobina c, sabatini a, deiana am, 2014. karyotype, ribosomal genes, and telomeric sequences in the crayfish procambarus clarkii (decapoda: cambaridae). j. crust. biol. 34:525-531. schmidt m, mellon d, 2011. neuronal processing of chemical information in crustaceans, p. 123-147. in: t. breithaupt and m. thiel (eds.), chemical communication in crustaceans. springer, new york. simon jl, moore pa, 2007. male-female communication in the crayfish orconectes rusticus: the use of urinary signals in no nco mm er cia l u se on ly mating behaviour in the red swamp crayfish 73 reproductive and non-reproductive pairings. ethology 113:740-754. solari p, melis m, sollai g, masala c, palmas f, sabatini a, crnjar r, 2015. sensing with the legs: contribution of pereiopods in the detection of food-related compounds in the red swamp crayfish procambarus clarkii. j. crustac. biol. 35:81-87. solari p, sollai g, masala c, loy f, palmas f, sabatini a, crnjar r, 2017. antennular morphology and contribution of aesthetascs in the detection of food-related compounds in the shrimp palaemon adspersus rathke, 1837 (decapoda: palaemonidae). biol. bull. 232:110-122. solari p, peddio s, sollai g, masala c, podda c, frau g, palmas f, sabatini a, crnjar r, 2018. development of pvc dispensers for long-lasting release of attractants for the control of invasive crayfish populations. diversity 10, 128. souty-grosset c, anastácio mp, aquiloni l, banha f, choquer j, chucholl c, tricarico e, 2016. the red swamp crayfish procambarus clarkii in europe: impacts on aquatic ecosystems and human well-being. limnologica 58:78-93. stebbing pd, bentley mg, watson gj, 2003. mating behaviour and evidence for a female released courtship pheromone in the signal crayfish pacifastacus leniusculus. j. chem. ecol. 29:465-475. subramoniam t, 2016. sexual biology and reproduction in crustaceans. academic press, amsterdam: 526 pp. thorp jh, 1984. theory and practice in crayfish communication studies. j. chem. ecol. 10:1283-1287. tierney aj, thompson cs, dunham dw, 1984. site of pheromone reception in the crayfish orconectes propinquus (decapoda, cambaride). j. crust. biol. 4:554-559. villanelli f, gherardi f, 1998. breeding in the crayfish, austropotamobius pallipes: mating patterns, mate choice and intermale competition. j. fresh. biol. 40:305-315. wyatt td, 2014. pheromones and animal behavior: chemical signals and signatures. cambridge university press, cambridge: 419 pp. no nco mm er cia l u se on ly layout 1 introduction the anthropogenic impact is significantly altering aquatic ecosystems (vörösmarty et al., 2010; elliott and elliott, 2013), by means of a plethora of pressures that include chemical pollution and wastewater discharge, eutrophication, hypoxia, utilization of living resources, habitat destruction and climate change effects (halpern et al., 2008). aquatic sediments, acting as repositories of materials and solutes deposited from or diffused through the water column, accumulate chemical and biological contaminants (ridgway and shimmield, 2002). this holds true also for microbiological pollutants, including autochthonous and pathogenic microbes of fecal origin, that reach the aquatic environment by a variety of routes and can spread diseases to human and aquatic populations, with important socio-economic, sanitary and environmental consequences (stewart et al., 2008). a number of studies have so far investigated the presence and distribution of fecal bacteria in aquatic sediments. the majority of these studies have addressed the traditional fecal indicators, such as total coliforms, escherichia coli and intestinal enterococci, which are worldwide used for assessments of aquatic ecosystems quality (field and samadpour, 2007 and references therein; liang et al., 2015). these studies have shown that substantial populations of fecal bacteria can be often retrieved in lagoon, estuarine and coastal sediments (an et al., 2002; luna et al., 2010; pachepsky and shelton, 2011; perini et al., 2015), suggesting that sediments are environmental reservoirs of fecal bacteria. however, the reliability of these indicators has been recently questioned, as they can persist and regrow in the environment, they are recovered also in absence of obvious fecal sources, and following the discovery of environmentally-adapted populations of e. coli (luo et al., 2011; byappanahalli et al., 2012). altogether, these issues have stimulated new studies, aimed at identifying alternative and more reliable indicators of fecal pollution, able to identify risks to human health and improve monitoring strategies (stewart et al., 2008). kreader (1995) pioneered the use of fecal anaerobes within the genus bacteroides as more reliable fecal indicators, and highlighted their potential to distinguish human from non-human sources of pollution. since then, several studies have been performed to discover and test new indicators of fecal pollution in water (reviewed in advances in oceanography and limnology, 2016; 7(2): 115-124 article doi: 10.4081/aiol.2016.5948 this work is licensed under a creative commons attribution-noncommercial 4.0 international license (cc by-nc 4.0). next generation sequencing reveals distinct fecal pollution signatures in aquatic sediments across gradients of anthropogenic influence gian marco luna,1 grazia m. quero,2 laura perini2 1cnr-ismar institute of marine sciences, national research council, largo fiera della pesca 2, 60125 ancona, italy; 2cnr-ismar institute of marine sciences, national research council, arsenale tesa 104, 30122 venezia, italy *corresponding author: gianmarco.luna@ismar.cnr.it abstract aquatic sediments are the repository of a variety of anthropogenic pollutants, including bacteria of fecal origin, that reach the aquatic environment from a variety of sources. although fecal bacteria can survive for long periods of time in aquatic sediments, the microbiological quality of sediments is almost entirely neglected when performing quality assessments of aquatic ecosystems. here we investigated the relative abundance, patterns and diversity of fecal bacterial populations in two coastal areas in the northern adriatic sea (italy): the po river prodelta (prp, an estuarine area receiving significant contaminant discharge from one of the largest european rivers) and the lagoon of venice (lv, a transitional environment impacted by a multitude of anthropogenic stressors). from both areas, several indicators of fecal and sewage contamination were determined in the sediments using next generation sequencing (ngs) of 16s rdna amplicons. at both areas, fecal contamination was high, with fecal bacteria accounting for up to 3.96% and 1.12% of the sediment bacterial assemblages in prp and lv, respectively. the magnitude of the fecal signature was highest in the prp site, highlighting the major role of the po river in spreading microbial contaminants into the adjacent coastal area. in the lv site, fecal pollution was highest in the urban area, and almost disappeared when moving to the open sea. our analysis revealed a large number of fecal operational taxonomic units (otu, 960 and 181 in prp and lv, respectively) and showed a different fecal signature in the two areas, suggesting a diverse contribution of human and non-human sources of contamination. these results highlight the potential of ngs techniques to gain insights into the origin and fate of different fecal bacteria populations in aquatic sediments. key words: 16s rdna; fecal bacteria; aquatic sediments; sewage; lagoon. received: april 2016. accepted: june 2016. no nco mm er cia l u se on ly 116 next generation sequencing of fecal bacteria mclellan and eren, 2014), while studies of alternative fecal indicators in aquatic sediments have been rare (kim and wuertz, 2015). on the light of the recognized role of aquatic sediments as reservoir of fecal bacteria, the urgency of expanding our poor knowledge in the sedimentary environment becomes manifestly evident. recent studies have shown the usefulness of next generation sequencing (ngs) technologies in water quality assessments (vierheilig et al., 2015). ngs techniques can provide insights into the ecology of microbe-mediated processes (such as biodegradation of contaminants and algal blooms) influencing water quality, but can also be extremely useful to identify an array of other taxa that could serve as indicators of fecal contamination in aquatic environments (tan et al., 2015). this holds true for those ngs methods that target small subunit rrna hypervariable regions, which are able to resolve microbial community composition in environmental samples, and to provide information on source-specific phylotypes and/or assemblages of phylotypes (newton et al., 2011, 2013). fisher et al. (2015) have recently shown the potential of sequencing the v6 region of the 16s rrna genes to discern between human versus non-human fecal sources. cumulatively, these emerging studies show the potential of ngs as powerful tool in detecting alternative fecal indicators also in aquatic sediments. we report here the results of an investigation, carried out in two different aquatic environments in the northern adriatic sea (italy), with the aim of assessing the presence and spatial distribution of traditional and alternative (feces-associated and sewer infrastructure-associated) fecal indicators by using ngs methods targeting the 16s rrna bacterial gene. this study is, to the best of our knowledge, among the first performed so far to explore the presence and spatial variability of alternative fecal indicators in different aquatic sediments, providing potentially useful insights for tracking the source and fate of fecal bacteria in aquatic sediments. methods description of the study areas sediments were collected in two coastal areas located in the northern adriatic sea: the po river prodelta (hereafter defined prp) and the lagoon of venice (hereafter defined lv), supposed to be exposed to different types of fecal contamination. the prp receives a significant discharge of contaminants from one of the largest rivers in europe (boldrin et al., 2005). the river discharges a mean of 1500 m3 s–1 of freshwater (with peaks up to >10,000 m3 s–1 during floods), that produces a freshwater plume able to influence the microbial diversity and functioning of the adjacent coastal ecosystem (manini et al., 2004, quero et al., 2015). the po experiences typically major floods, which transport large amounts of suspended sediments and associated pollutants to the sea, that can be stored or transported offshore to the adjacent marine areas (correggiari et al., 2005). many studies have shown that the coastal area in front of the delta is severely chemically polluted, as the river carries yearly tons of anthropogenic chemicals collected from the po valley and the river tributaries, including emerging contaminants (casatta et al., 2015). the lv, among the largest lagoons in the mediterranean, is a microtidal, semi-closed lagoon connected to the adriatic sea by three openings (inlets), influenced by a tidal semidiurnal regime (cucco and umgiesser, 2006). the tidal regime governs the water exchange with the adjacent open sea, and largely influences the renewal capacity of the lagoon basin. the lv is historically impacted by a multitude of anthropogenic stressors, among which one of the largest industrial plants of italy located on the nearby mainland, touristic and commercial ports, several small tributaries (that carry around 30-35 m3 s–1 of freshwater; zuliani et al., 2005), agricultural and municipal wastes. as far as the fecal contamination is concerned, an important source of contamination is the city of venice which, due to its history and unique building architecture, has never been provided with a modern and efficient sewage treatment system (sfriso and facca, 2013). consequently, only partially treated effluents from a large number of domestic and commercial inputs are discharged daily into the city canals, determining a diffuse contamination and the accumulation of fecal bacteria in sediments and live macroalgae (quero et al., 2015; perini et al., 2015). sampling activities in the prp area, sampling was performed between 10th and 14th june 2013 in front of the outlets of the main branches of the river delta. the sampling design included 11 stations (fig. 1), distributed along coast-to-open sea transects, at depths comprised between 9.5 and 20.5 meters. the geographic coordinates (as longitude and latitude) and water depths of the stations were as follows: 1 (12.543 e, 44.99 n, 12 m), 2 (12.556 e, 44.993 n, 20.5 m), 3 (12.571 e, 44.968 n, 9.5 m), 4 (12.576 e, 44.971 n, 15 m), 5 (12.581 e, 44.974 n, 19 m), 6 (12.573 e, 44.955 n, 10.5 m), 7 (12.585 e, 44.952 n, 14 m), 8 (12.558 e, 44.929 n, 11 m), 9 (12.573 e, 44.919 n, 14.6 m), 10 (12.538 e, 44.891 n, 14.8 m) and 11 (12.558 e, 44.886 n, 17.5 m). the sampling transects were in front of the outlets of the main branches of the river delta. the spatial distribution of the stations was thought to follow the possible deposition route of pollutants transported by an exceptional flood event, like the one that occurred in the third week of may 2013, with a maximum flow rate of 6830 m3 s–1. more details about the sampling activities and the characteristics of the sampling stations in the prp are reported in quero et al. (2015). no nco mm er cia l u se on ly 117g.m. luna et al. in the lv area, sampling was performed during the autumn season (21st-22th october 2014), chosen as the period of highest river runoff, expected to increase the load of fecal bacteria within the lagoon. the sampling design included 5 sampling stations (fig. 1), at depths comprised between 5.4 and 16 m, distributed across a gradient of putative contamination from the inner part of the lagoon to the open sea. the geographic coordinates (longitude and latitude) and water depth of the stations were as follows: industrial port (12.219 e, 45.438 n, 5.4 m), inner lagoon (12.258 e, 45.448 n, 7.2 m), cruise port (12.311 e, 45.436 n, 12.1 m), city centre (12.352 e, 45.431 n, 7.1 m) and open sea (12.508 e, 45.313 n, 16 m). in both areas, sediments were collected using a van veen grab sampler (sampling surface 0.1 m2), onboard small research vessels (a privately operated one in the prp area, and the ‘litus’ boat operated by ismar-cnr in the lv one). once onboard, the uppermost 0-2 cm layer of sediment was immediately placed, using sterile spatulas, in sterile containers for their immediate transport at 4°c to the laboratory, where the samples were stored at -20°c until analysis. analyses of fecal bacteria using illumina sequencing of the 16s rrna gene dna was extracted from one gram of each sediment sample using the powersoil® dna isolation kit (mobio fig. 1. the two study sites in the northern adriatic sea (italy), with indication and name of the sampling stations (blue dots). latitude (n) and longitude (e) are reported; lv, lagoon of venice; prp, po river prodelta. no nco mm er cia l u se on ly 118 next generation sequencing of fecal bacteria laboratories inc., california), according to the manufacturer’s instructions with some slight modifications to increase the dna yield and quality. these modifications included two additional vortexing steps (following the one which is recommended by the manufacturer) at the maximum speed for 2 min, each one being preceded by an incubation at 70°c for 5 min, and by one more washing step with solution c5 as an additional removal step for contaminants. the concentration of each dna extract was determined spectrophotometrically using nanodrop ( thermo scientific) and the dna was stored at -80°c until pcr. illumina miseq v3 sequencing were carried out on the hypervariable v3 and v4 regions of the 16s rrna gene by amplifying using the 341f (5′−cctacgggnggcwgcag−3′) and 785r (5′−gactachvgggtatctaatcc−3′) universal bacterial primers (eiler et al., 2012). paired-end reads were quality checked (with default settings and minimum quality score of 20) and analyzed with qiime v1.8.0 software package (quantitative insights into microbial ecology; caporaso et al., 2010). reads were clustered into otus by using uclust v1.2.22 (edgar, 2010) with a >97% similarity threshold with an open-reference otu picking strategy and default settings. chimeras were detected by using usearch v6.1 (edgar, 2010). chimera checking and taxonomy assignment was performed using greengenes 13.8 as reference database (desantis et al., 2006). to account for differences in the sequencing effort among samples, abundances in each sample were normalized to the number of sequences of the sample showing the lowest number of reads. the sequences of the prp area have been submitted to the sra (sequence read archive; accession numbers srp061637), while those for the lv are currently being submitted. data elaboration and statistical analyses within each of the bacterial assemblage, we searched for those otus identified as belonging to the traditional fecal indicator taxa (i.e., the family enterobacteriaceae, that includes the genus escherichia, and the genus enterococcus), and those otus belonging to the alternative fecal indicator taxa, according to the approach proposed by newton et al. (2013). as alternative indicators, we searched for otus belonging to five feces-associated bacterial families (bacteroidaceae, porphyromonadaceae, clostridiaceae, lachnospiraceae and ruminococcaceae) and three sewer infrastructure-associated bacterial genera (acinetobacter, arcobacter and trichococcus), that we used here as signatures of fecal (human and non-human) and sewage contamination, respectively. this distinction was based on the study by newton et al. (2013), who reported that five feces-associated bacterial families were prevalent (up to 85% of the total sequences) in the feces of animals and humans while, on the other hand, three sewer infrastructure-associated genera were very abundant in the sewage samples while not prevalent in human feces (only 33 sequences recovered out of a >1.2 million sequences of a human fecal dataset). vandewalle et al. (2012) reported that only a small fraction of otus in sewage matched sequences from human fecal samples, and suggested that these sewage-associated taxa, that thrive within the sewer system, may serve as useful adjuncts to fecal indicators for tracking sewage pollution in surface waters. the spearman-rank correlation analysis was performed to test for relationships between the relative abundance of traditional and alternative fecal indicators. correlation coefficients (r) were considered significant at p-values less than 0.05. differences in the composition of fecal bacterial communities between the sampling areas, and between groups of stations within each of the two areas, were assessed by using the analysis of similarity (anosim) tool based on a bray-curtis similarity matrix. the presence of statistical differences between samples is indicated by a significance level at p-values less than 0.05. to assess differences in the composition of indicator otus between the sampling areas, we applied univariate distance-based permutational analyses of variance (permanova). the statistical analyses were carried out using a sampling design, that considered the area as fixed factor as source of variance, with 2 levels (prp and lv). the anosim and permanova analyses were performed using the primer 6 software (http://www.primer-e.com/). results relative abundance of traditional, fecesand sewage-associated indicators at both areas, sequence analyses of the sediment samples revealed evidences for a marked and diffuse fecal signature that was, however, significantly different between the two areas (anosim, r=0.314, p<0.05). in the prp area, the relative abundance of traditional indicators (fig. 2a) accounted for 0.01 to 0.19% of the bacterial assemblage in the case of enterobacteriaceae, and for 0 (no sequences detected) to 0.01% in the case of enterococcus. in the same study area, the relative abundance of alternative fecal indicators was quite higher than that of traditional indicators. the contribution of the feces-associated indicators (fig. 2c) was in the range 0.08-2.13% (lachnospiraceae), 0.17-0.53% (clostridiaceae), 0-0.14% (porphyromonadaceae), 0-0.04% (bacteroidaceae) and 0.08-0.63% (ruminococcaceae), while the contribution of the sewage-associated indicators (fig. 2e) was in the range 0-0.17% (arcobacter), 0.06-0.57% (acinetobacter), with only one or two sequences per sample assigned to the trichococcus genus (only at the stations 4, 5, 6 and no nco mm er cia l u se on ly 119g.m. luna et al. 8). the cumulative contribution of all the sequences belonging to the traditional, fecesand sewage-associated indicators was in the range 0.49-3.96% of the bacterial assemblages (on average 1.63%). in the lv area, the level of fecal contamination was overall reduced when compared to the prp area. as far as the traditional indicators are concerned, these were observed in the lv area at all stations (fig. 2b), in the range 0-0.05% (enterobacteriaceae) and 0.004-0.03% (enterococcus). the relative abundance of the feces-associated indicators (fig. 2d) was higher and accounted for 0.050.40% (lachnospiraceae), 0.01-0.29% (clostridiaceae), 0-0.02% (porphyromonadaceae), 0-0.02% (bacteroidaceae) and 0.001-0.27% (ruminococcaceae), while the contribution of the sewage-associated indicators (fig. 2f) was in the range 0.01-0.11% (arcobacter), 00.06% (acinetobacter), with no sequences assigned to the genus trichococcus. the cumulative contribution of all of the sequences belonging to the traditional and alternative indicators accounted from 0.08 to 1.12% of the bacterial assemblage (on average 0.60%). spatial patterns of fecal indicators in the two areas in the prp area, the fecal contamination in stations located closer to the coast (stations 1, 3, 4, 5, 6, 8 and 10; average relative abundance 2.0%) was two-fold higher that in offshore stations (2, 7, 9 and 11; average 0.99%). the highest abundance of fecal indicators as a whole (i.e., summing up traditional and alternative indicators) was observed in the stations located right in front of the main outlet of the po river (stations 4, 5, and 6; up to 3.96% in the station 6). the relative abundance of traditional indicators didn’t show significant correlation either with the fecesor with the sewage-associated indicators (p≥0.05 for both relationships), whereas the fecesand sewage-associated indicators were significantly and positively related (r=0.83, p<0.01). in the lv area, the fecal contamination was much higher inside the lagoon (stations ind. port, inner lagoon, cruise port and city centre; average relative abundance 0.73%), and only weakly detectable in the open sea station (relative abundance 0.08%). within the lagoon, the highest abundance of fecal indicators as a whole was observed in the station located closer to the city center of venice (1.12%), followed by the two stations located close to the industrial area (station ind. port, 0.64%) and the mainland (inner lagoon, 0.70%). in the lv area the different types of fecal indicators were positively and significantly correlated (r=0.82, p<0.05 between traditional and feces-associated indicators; r=0.92, p<0.01 between traditional and sewage-associated indicators). the fecesand sewage-associated indicators were also significantly and positively correlated (r=0.751, p<0.05). traditional, fecal and sewage indicator otus the number of otus that were affiliated with each of the fecal indicator groups varied widely between areas and among stations in each area, ranging from 0 to 266 fig. 2. bar plots of the relative abundance of 16s rdna sequences belonging to: a, b) the traditional fecal indicator bacterial taxa (enterobacteriaceae and enterococcus); c, d) the feces-associated bacterial families (lachnospiraceae, clostridiaceae, porphyromonadaceae, bacteroidaceae and ruminococcaceae); and e, f) the sewage-associated bacterial genera (arcobacter, acinetobacter and trichococcus) in the sediments of the two sites. in the lv site: ip, industrial port; il, inner lagoon; cp, cruise port; cc, city centre; os, open sea. no nco mm er cia l u se on ly 120 next generation sequencing of fecal bacteria otus depending on the fecal indicator (tab. 1). the results of univariate permanova revealed a significant effect of the factor area (p<0.01). in the prp area, the results are summarized by grouping the stations located closer to the coast (1, 3, 4, 5, 6, 8 and 10) and those located more offshore (stations 2, 7, 9 and 11). this choice was supported by the anosim analysis, that demonstrated significant differences, in terms of otu community composition, between these two groups of stations (anosim, p<0.05). in this area, the cumulative number of otus (i.e., the sum of all otus recorded at all stations) belonging to the traditional indicators was 46 for enterobacteriaceae and 4 for enterococcus (only observed in the more coastal stations). the number of otus within the feces-associated indicators was much higher: 176 and 98 for lachnospiraceae (coastal and offshore stations, respectively), 169 and 141 (clostridiaceae), 55 and 36 (porphyromonadaceae), 22 and 9 (bacteroidaceae), 132 and 93 (ruminococcaceae). the number of otus associated with the sewage-associated indicators was 3 and 0 for trichococcus (coastal and offshore stations, respectively), 69 and 54 (acinetobacter), 19 and 34 (arcobacter). overall, in the prp area the number of otus in the coastal stations was higher than in the offshore ones, with only one exception (i.e., arcobacter otus more abundant in offshore than in coastal stations). in lv area, the cumulative number of otus belonging to the traditional indicators was 5 and 2 (for enterobacteriaceae and enterococcus, respectively). the cumulative number of the fecal-associated indicator otus was larger compared to the traditional ones, corresponding to 38 (lachnospiraceae, range among sampling stations 4-18), 37 (clostridiaceae, range 5-18), 16 (porphyromonadaceae, range 0-10), 18 (bacteroidaceae, range 0-9) and 22 otus (ruminococcaceae, range 0-10). at the same time, the number of otus belonging to the sewage-associated indicators was 12 for acinetobacter (range 0-5) and 31 for arcobacter (range 4-20), with no otus belonging to trichococcus. the highest number of otus was observed in the lagoon stations (namely, at stations ind. port, inner lagoon and city centre. the station located in the open sea showed only 14 otus for the entire pool of fecal indicators (traditional, fecesand sewage-associated). discussion in this study, we aimed at examining the presence, prevalence and spatial distribution of traditional and alternative fecal indicators in a range of aquatic sediments, collected in transitional, estuarine and coastal marine areas along gradients of anthropogenic influence. we used ngs of 16s rrna gene amplicons to identify and track fecal indicator bacteria within complex benthic microbial assemblages, by taking advantage of the method recently proposed by newton et al. (2013), that allows to potentially discriminate between feces-associated and sewageassociated bacteria. the presence and distribution patterns of alternative indicators in the sediments were then compared with those of traditional indicators (escherichia coli, within the enterobacteriaceae family, and enterococci) that are utilized worldwide to assess fecal pollution tab. 1. number of otus affiliated with the different fecal indicator bacteria in the two study sites. for the po river prodelta site, results are summarized by summing all the otus observed at stations 1, 3, 4, 5, 6, 8 and 10 (stations close to the coast) and those observed at stations 2, 7, 9 and 11 (offshore stations). prp lv microbial number of otus number of otus cumulative indicator stations close offshore cumulative stations within the lagoon no. of otus to the coast stations no. of otus ip il cp cc open sea traditional enterobacteriaceae 27 28 46 1 5 1 3 0 5 enterococcus 4 0 4 1 1 1 3 1 2 feces-associated lachnospiraceae 176 98 266 17 18 9 9 4 38 clostridiaceae 169 141 231 10 16 18 10 5 37 porphyromonadaceae 55 36 70 3 7 2 10 0 16 bacteroidaceae 22 9 28 3 9 4 8 0 18 ruminococcaceae 132 93 187 6 10 10 6 0 22 sewage-associated trichococcus 3 0 3 0 0 0 0 0 0 acinetobacter 69 54 85 5 4 1 3 0 12 arcobacter 19 34 40 20 9 8 17 4 31 prp, po river prodelta; lv, lagoon of venice; otu, operational taxonomic units; ip, industrial port; il, inner lagoon; cp, cruise port; cc, city centre; os, open sea. no nco mm er cia l u se on ly 121g.m. luna et al. in aquatic environments. recently, a large body of studies have identified and tested alternative indicators of fecal pollution in aquatic systems, by exploiting the ngs technologies to provide unprecedented inventories of microbial communities in aquatic samples (mclellan and eren, 2014). however, while a large number of studies have investigated the presence and spatial patterns of alternative indicators in water (savichtcheva and okabe, 2006; liang et al., 2015), similar studies in sediments have been rare (kim and wuertz, 2015). this lack of knowledge is surprising, given the recognized role of aquatic sediments as reservoir of fecal bacteria (luna et al., 2010), and their potential to favor the persistence of fecal microbes and to contaminate the overlying water through resuspension, which likely poses important public health and environmental threats. at both study areas, we found evidences for an extensive microbial pollution that was testified, despite at different extent, by the presence of traditional, fecesand sewage-associated indicators. these indicators accounted cumulatively for a variable fraction of benthic assemblages with peaks, as in the case of the most polluted sediments, up to 3.96%. at both study areas, the magnitude of the fecal signature decreased with increasing distance from the sources of pollution, which underlines the usefulness of this approach, that includes also the alternative indicators, to track the fecal pollution in aquatic sediments. the presence of fecal pollution in the two sites was not unexpected, since they are both subjected to a significant anthropogenic pressure, and likely receive important loads of fecal bacteria. the prp coastal area receives a significant discharge of sediments and pollutants from the largest italian river, especially after the floods that occur on a seasonal (biannual) frequency and, also, as episodic short-term events (palinkas et al., 2005). its drainage basin contributes to more than 35% of the national agricultural, livestock and industrial production, that originates organic loads estimated in 114×106 inhabitant equivalents (casatta et al., 2015). however, while the presence of chemical pollutants in the coastal area near the mouth of the po river has been largely reported (romano et al., 2013), this is the first report on the presence of extensive fecal pollution in these sediments. the largest fecal contamination was observed in the stations located in front of mouth of the pila distributary, which is the main outlet of the river and can discharge up to 70% of the sediment load delivered to the sea. however, the fecal signature was also observed in stations located downstream of this main mouth, that suggests either a fecal input from the other minor distributaries, and/or the potential of fecal bacteria of being distributed over a large coastal area. this bacterial spread may occur when the high fluvial discharge coincides with energetic physical oceanographic conditions, preventing deposition in shallow waters and favoring sediment dispersion in seabed deposits located offshore and downstream of the mouth (palinkas et al., 2005). our finding of a large reservoir of fecal bacteria in the po prodelta sediments poses a claim for potential health consequences, given the potential of sediment bacteria of being re-suspended and transported southward, where several bathing beaches and touristic destinations are present. nevertheless we point out here that further studies are needed to investigate the ability of the different fecal bacterial populations, including the fecesand sewage-associated bacteria, to persist or decay once they reach the marine environment, and of being potentially transported toward adjacent coastal areas in presence of specific hydrological conditions. the finding of a diffuse fecal contamination in the lv area, especially in the stations more exposed to anthropogenic impacts, likely depends on the wide variety of human impacts, in the form of large industrial plants, touristic and commercial ports, rivers, agricultural and municipal wastes, that have affected this vulnerable transitional environment in the last years (micheletti et al., 2011). our results show that the highest fecal contamination is observed in the area closer to the city of venice, confirming recent findings of a chronic fecal pollution (perini et al., 2015), but also that other areas of the venice lagoon, located closer to the mainland, receive important loads of bacteria of fecal origin. this may be due to multiple delivery routes discharging in the inner part of the lagoon, that include runoff, outfall discharge, sewer overflows (that causes untreated sewage to be released), presence of tributaries, and other human activities. it is worth noticing that a signal, despite weak, of fecal contamination was observed also in the sediments of the open sea station. we speculate that the presence of fecal pollutants in this area, which is relatively far by obvious sources of pollution, could be the consequence of fecal discharges by offshore point sources and/or by the underwater submarine wastewater pipes present in the area (scroccaro et al., 2010), that may disperse microbial pollutants over vast marine areas. we report here that different aquatic sediments subjected to anthropogenic influence contain important proportions of traditional indicator bacteria, as well as fecal bacteria representing signatures of fecal (human and animal) and sewage sources. at both areas, the relative abundance of fecaland sewage-associated bacteria always exceeded that of the traditional indicators. sediments are a potentially favorable environment for fecal microorganisms survival, and this may be especially true for many of the alternative indicators (such as bacteroidales), that do not survive well in water due to their obligate anaerobic physiology (bae and wuertz, 2015) but may persist longer in low-oxygen or anoxic layers that are common in sediments. in both areas, we found that the fecal signature was more evident than the sewage signature. however, the avno nco mm er cia l u se on ly 122 next generation sequencing of fecal bacteria erage ratio of the fecal pollution signature to the sewage pollution signature was different in the two areas. in prp, the average contribution of the feces-associated taxa was 1.244%, a percentage 3.9-fold higher than that of the sewage-associated taxa (0.320%). conversely, in the lv area, the average contribution of feces-associated taxa was 0.602% (excluding the open sea station), a percentage 7.05-fold higher than that of the sewage-associated indicators (0.085%). the prevalence of different fecal bacterial signatures in the two areas suggests the presence of different fecal sources. the higher ratio of fecesto sewage-associated bacteria observed in the lagoon of venice suggests that human/animal sources of pollution are more important than sewage in polluting this area while, in the po prodelta area, the sewage pollution appears to be an important source. the sewage pollution in the prp area is particularly evident in the acinetobacter signature which is, on average, 14-fold more abundant than in the lv site. this genus is the most important in sewage (newton et al., 2013), and its recovery in sediments suggests that it may be a reliable signature of sewage pollution also in the benthic environment. it remains to be determined whether members of this genus are adapted to survive in aquatic sediments, and how long are their decay rates once they reach the sedimentary environment. the importance of sewage in contributing to pollution in the po prodelta is also confirmed by the recovery of another sewage-associated genus (trichococcus) that, though with only a few sequences per sample, was observed only in this area and not in the venice lagoon. a very large number of fecal otus, accounting for a total of 960 and 181 otus (in the prp and lv areas, respectively), were observed in the sediments under scrutiny. the largest number of otus was observed in the po prodelta area. this may suggest the existence, in this area, of multiple delivery mechanisms containing multiple fecal sources, that likely originate from the large drainage basin, and resulting in more complex fecal otu signatures than those observed in the lv sediments. in many cases and at both areas, fecal bacteria populations were dominated by few dominant otus. a closer examination of the top most abundant fecal otus in the two sites provided additional insights, useful to potentially distinguish among different sources of pollution. in the prp area, the traditional indicators showed dominance of one enterobacteriaceae otu (range 5-177 sequences per sediment sample) that showed 100% blast identity with escherichia coli strain 732 (accession number cp015138). the enterococci populations, that were poorly represented at this area, were dominated by one otu affiliated with enterococcus casseliflavus isolated from cow rumen (accession number kt630829), suggesting that cattle may be an important route of fecal pollution in this area. conversely, in the lv area, the dominant enterobacteriaceae otu was affiliated with yersinia kristensenii (accession number hg938308.1) while the dominant enterococcal otu, especially abundant in the station closer to the city center, showed a top blast match with catellicoccus marimammalium (accession number kf251005.1) isolated from gulls’ feces. in the venice urban area, populations of urban gulls have increased exponentially in the last years (rock, 2012). our findings indicate that they can contribute as an additional source of fecal contamination in the lagoon, which deserves further investigations. as far as the fecaland sewage-associated otus are concerned, the same analysis of the identity of the dominant otus revealed also interesting additional insights. the dominant acinetobacter otus in the lv area, particularly abundant in the station closer to the city center, showed a top blast match with an uncultured acinetobacter (hq742373.1) associated with the human intestine, confirming human fecal pollution as an important contamination route in the city (perini et al., 2015). conversely, the analyses of some of the dominant lachnospiraceae otus showed, at both prp and lv areas, top blast identities with several uncultured bacteria, which were observed in a variety of sediments and also in coastal sediments vegetated by seagrasses (jensen et al., 2007), apparently far from fecal pollution sources. these preliminary findings suggest that some of members within this family may be part of the natural benthic assemblages, and are thus not reliable indicators of fecal pollution. it is evident that, given also the much higher complexity of benthic microbial assemblages when compared with the planktonic ones, there is still much to be deciphered when using this type of community approaches to track fecal pollution in aquatic sediments. overall, our results demonstrate that the coupled analyses of the diversity and magnitude of the three fecal signatures (traditional, fecaland sewage-associated bacteria) are useful to discern among different pollution sources in the sediments of transitional, estuarine and coastal areas. conclusions we have shown that a wide range of aquatic sediments in areas exposed to anthropogenic stressors host important proportions of traditional indicator bacteria, but also of alternative bacterial taxa that more specifically track the presence of fecal (human and animal) and sewage pollution. the magnitude and pattern of the complex fecal signature followed the expected gradients of microbial pollution, and provided information useful to identify the main sources of pollution in the two study sites. our results also emphasize the opportunities that ngs techniques now offer to disentangle complex fecal pollution signals, and to source track and identify alternative fecal indicators in lagoon and marine sediments. no nco mm er cia l u se on ly 123g.m. luna et al. acknowledgments this work was possible thanks to funds granted to gml by the programme ritmare (sp3-wp2-a2 strumenti innovativi per la valutazione degli effetti di contaminanti emergenti sulle comunità biologiche), the ipa project balmas ballast water management system for adriatic sea protection (project code 1° str/0005) funded by eu, the short term mobility (stm) program funded by the cnr, and partially to funds granted to gmq by the malware metagenomic study of microbial invasive species introduced by ballast waters project (funded by lifewatch italia). we sincerely thank dr. francesca garaventa (ismar-cnr) for her help in sampling the prp site, and similarly the crew of the research boat ‘litus’ and dr. mauro bastianini (ismar-cnr) for making sample collection possible in the lv site. references an yj, kampbell dh, breidenbach gp, 2002. escherichia coli and total coliforms in water and sediments at lake marinas. environ. pollut. 120:771-778. bae s, wuertz s, 2015. decay of host-associated bacteroidales cells and dna in continuous-flow freshwater and seawater microcosms of identical experimental design and temperature as measured by pma-qpcr and qpcr. wat. res. 70: 205-213. byappanahalli mn, nevers mb, korajkic a, staley zr, harwood vj, 2012. enterococci in the environment. microbiol. mol. biol. rev. 76:685-706. boldrin a, langone l, miserocchi s, turchetto m, acri f, 2005. po river plume on the adriatic continental shelf: dispersion and sedimentation of dissolved and suspended matter during different river discharge rates. mar. geol. 222:135-158. caporaso jg, kuczynski j, stombaugh j, bittinger k, bushman fd, costello ek, fierer n, pena ag, goodrich jk, gordon ji, huttley ga, kelley st, knights d, koenig je, ley re, lozupone ca, mcdonald d, muegge bd, pirrung m, reeder j, sevinsky jr, turnbaugh pj, walters wa, widmann j, yatsunenko t, zaneveld j, knight r, 2010. qiime allows analysis of high-throughput community sequencing data. nat. methods 7:335-336. casatta n, stefani f, pozzoni f, guzzella l, marziali l, mascolo g, viganò l, 2015. endocrine-disrupting chemicals in coastal lagoons of the po river delta: sediment contamination, bioaccumulation and effects on manila clams. environ. sci. pollut. res. int. 23:10477-10493. correggiari a, cattaneo a, trincardi f, 2005. the modern po delta system: lobe switching and asymmetric prodelta growth. mar. geol. 222:49-74. cucco a, umgiesser g, 2006. modeling the venice lagoon residence time. ecol. model. 193:34-51. desantis tz, hugenholtz p, larsen n, rojas m, brodie el, keller k, et al., 2006. greengenes, a chimera-checked 16s rrna gene database and workbench compatible with arb. appl. environ. microbiol. 72:5069-5072. edgar rc, 2010. search and clustering orders of magnitude faster than blast. bioinformatics 26:2460-2461. eiler a, heinrich f, bertilsson s, 2012. coherent dynamics and association networks among lake bacterioplankton taxa. isme j. 6:330-342. elliott je, elliott kh, 2013. tracking marine pollution. science 340: 556-558. field kg, samadpour m, 2007. fecal source tracking, the indicator paradigm, and managing water quality. water res. 41:3517-3538. fisher jc, eren am, green hc, shanks oc, morrison hg, vineis jh, sogin ml, mclellan sl, 2015. comparison of sewage and animal fecal microbiomes by using oligotyping reveals potential human fecal indicators in multiple taxonomic groups. appl. environ. microbiol. 81:7023-7033. halpern bs, walbridge s, selkoe ka, kappel cv, micheli f, d’agrosa c, et al., 2008. a global map of human impact on marine ecosystems. science 319:948-952. jensen si, kühl m, priemé a, 2007. different bacterial communities associated with the roots and bulk sediment of the seagrass zostera marina. fems microbiol. ecol. 62:108-117. kim m, wuertz s, 2015. survival and persistence of host-associated bacteroidales cells and dna in comparison with escherichia coli and enterococcus in freshwater sediments as quantified by pma-qpcr and qpcr. water res. 87:182-192. kreader ca, 1995. design and evaluation of bacteroides dna probes for the specific detection of human fecal pollution. appl. environ. microbiol. 61: 1171-1179. liang l, goh sg, vergara ggrv, fang hm, rezaeinejad s, chang sy, bayen s, lee wa, sobsey md, rose jb, gin kyh, 2015. alternative fecal indicators and their empirical relationships with enteric viruses, salmonella enterica, and pseudomonas aeruginosa in surface waters of a tropical urban catchment. appl. environ. microbiol. 81:850-860. luna gm, vignaroli c, rinaldi c, pusceddu a, nicoletti l, gabellini m, danovaro r, biavasco f, 2010. extraintestinal escherichia coli carrying virulence genes in coastal marine sediments. appl. environ. microbiol. 76:5659-5668. luo c, walk st, gordon dm, feldgarden m, tiedje jm, konstantinidis kt, 2011. genome sequencing of environmental escherichia coli expands understanding of the ecology and speciation of the model bacterial species. p. natl. acad. sci. usa 108:7200-7205. manini e, luna gm, danovaro r, 2004. benthic bacterial response to variable estuarine water inputs. fems microbiol. ecol. 50:185-194. mclellan sl, eren am, 2014. discovering new indicators of fecal pollution. trends microbiol. 22:697-706. micheletti s, gottardo a, critto a, chiarato a, marcomini a, 2011. environmental quality of transitional waters: the lagoon of venice case study. environ. int. 37:31-41. newton rj, bootsma mj, morrison hg, sogin ml, mclellan sl, 2013. a microbial signature approach to identify fecal pollution in the waters off an urbanized coast of lake michigan. microb. ecol. 65:1011-1023. newton rj, vandewalle jl, borchardt ma, gorelick mh, mclellan sl, 2011. lachnospiraceae and bacteroidales alternative fecal indicators reveal chronic human sewage contamination in an urban harbor. appl. environ. microbiol. 77: 6972-6981 no nco mm er cia l u se on ly 124 next generation sequencing of fecal bacteria pachepsky ya, shelton dr, 2011. escherichia coli and fecal coliforms in freshwater and estuarine sediments. crit. rev. environ. sci. technol. 41: 067-1110. palinkas cm, nittrouer ca, wheatcroft ra, langone l, 2005. the use of 7 be to identify event and seasonal sedimentation near the po river delta, adriatic sea. mar. geol. 222:95-112. perini l, quero gm, garcía es, luna gm, 2015. distribution of escherichia coli in a coastal lagoon (venice, italy): temporal patterns, genetic diversity and the role of tidal forcing. water res. 87:155-165. quero gm, cassin d, botter m, perini l, luna gm, 2015. patterns of benthic bacterial diversity in coastal areas contaminated by heavy metals, polycyclic aromatic hydrocarbons (pahs) and polychlorinated biphenyls (pcbs). front. microbiol. 6:1053. ridgway j, shimmield g, 2002. estuaries as repositories of historical contamination and their impact on shelf seas. est. coast. shelf sci. 55: 903-928. rock p, 2012. urban gulls. why current control methods always fail. riv. ital. ornitol. 82: 58-65. romano s, langone l, frignani m, albertazzi s, focaccia p, bellucci lg, ravaioli m, 2013. historical pattern and mass balance of trace metals in sediments of the northwestern adriatic sea shelf. mar. poll. bull. 76: 32-41. savichtcheva o, okabe s, 2006. alternative indicators of fecal pollution: relations with pathogens and conventional indicators, current methodologies for direct pathogen monitoring and future application perspectives. water res. 40: 2463-2476. scroccaro i, ostoich m, umgiesser g, de pascalis f, colugnati l, mattassi g, vazzoler m, cuomo m, 2010. submarine wastewater discharges: dispersion modelling in the northern adriatic sea. environ. sci. pollut. res. 17: 844-855. sfriso a, facca c, 2013. annual growth and environmental relationships of the invasive species sargassum muticum and undaria pinnatifida in the lagoon of venice. est. coast. shelf sci. 129:162-172. stewart jr, gast rj, fujioka rs, solo-gabriele hm, meschke js, amaral-zettler la, del castillo e, polz mf, collier tk, strom ms, sinigalliano cd, moeller pdr, 2008. the coastal environment and human health: microbial indicators, pathogens, sentinels and reservoirs. environ. health 7:s3. tan b, ng c, nshimyimana jp, loh ll, gin ky-h, thompson jr, 2015. next-generation sequencing (ngs) for assessment of microbial water quality: current progress, challenges, and future opportunities. front. microbiol. 6:1027. vandewalle jl, goetz gw, huse sm, morrison hg, sogin ml, hoffmann rg, yan k, mclellan sl, 2012. acinetobacter, aeromonas and trichococcus populations dominate the microbial community within urban sewer infrastructure. environ. microbiol. 14: 2538-2552. vierheilig j, savio d, ley re, mach rl, farnleitner ah, reischer gh, 2015. potential applications of next generation dna sequencing of 16s rrna gene amplicons in microbial water quality monitoring. wat sci technol 72:1962-1972. vörösmarty cj, mcintyre pb, gessner mo, dudgeon d, prusevich a, green p, glidden s, bunn se, sullivan ca, liermann cr, davies pm, 2010. global threats to human water security and river biodiversity. nature 67:555-561. zuliani a, zaggia l, collavini f, zonta r, 2005. freshwater discharge from the drainage basin to the venice lagoon (italy). environ. internat. 31:929-938. no nco mm er cia l u se on ly layout 1 introduction animals and plants can acclimatise to specific situations, adapting to environmental and biological changes; however, the impacts of the global climate change (gcc) exceed the capacity of many organisms to adapt due to their highly rapid onset (hughes et al., 2003; lejeusne et al., 2010; rossi, 2013). the increase of seawater temperature is one of the major effects of gcc; according to the most recent projections, a global temperature increase of 3°c is forecasted by the end of the current century (meinshausen et al., 2011; vuuren et al., 2008). in the nw mediterranean sea, major mortality events of marine rocky benthic communities, especially gorgonians, have occurred in 1999 and 2003 (cerrano et al., 2000; di camillo and cerrano, 2015; garrabou et al., 2001). such events have been directly or indirectly related to temperature increases that overcame natural seasonal fluctuations (ferrier-pagès et al., 2009). elevated temperatures reduce oxygen solubility (truesdale et al., 1955), thus limiting its availability to benthic suspension feeders. such reduced availability of oxygen coinciding with an increased metabolic oxygen demand may cause severe oxygen limitation in marine invertebrates under high temperatures (previati et al., 2010). detrimental effects related to mass mortality events (mmes) were observed to be ‘patchy’ for several species (cerrano et al., 2000), suggesting that resistance to such stressors could show an inter-individual variability. this findings highlights the need for improving our knowledge on the different responses to thermal stress within gorgonian species (crisci et al., 2011; ezzat et al., 2013; garrabou et al., 2009). mmes affecting the precious red coral corallium rubrum (l., 1758) have to be considered of great regard as this anthozoan, widely distributed in the mediterranean sea and neighbouring atlantic coasts, represent a key element within the so called ‘animal forests’ communities (cau et al., 2016, 2015; rossi, 2013; rossi et al., 2008). for c. rubrum, a range of mortality between 5% and 80% was recorded, depending on geographic locations and depth range of the affected populations, with partial to total mortality due to tissue necrosis (cerrano et al., 2000; di camillo and cerrano, 2015; garrabou et al., 2009, 2001; haguenauer et al., 2013). advances in oceanography and limnology, 2018; 9(1): 13-18 article doi: 10.4081/aiol.2018.7275 this work is licensed under a creative commons attribution-noncommercial 4.0 international license (cc by-nc 4.0). differential response to thermal stress of shallow and deep dwelling colonies of mediterranean red coral corallium rubrum (l., 1758) alessandro cau,1,2 lorenzo bramanti,3 rita cannas,1 davide moccia,1 bachisio mario padedda,2 cristina porcu,1 flavio sacco,1 maria cristina follesa1 1department of life and environmental sciences, university of cagliari, italy; 2department of architecture, design and urban planning, university of sassari, alghero, italy; 3cnrs, umr8222, lecob, oceanographic observatory, pierre and marie curie university, banyuls-sur-mer, france corresponding author: alessandrocau@unica.it abstract over the last decades, the global climate change caused increase in seawater temperature, which have shown to be detrimental for mediterranean red coral populations (corallium rubrum). recent researches on c. rubrum described how in shallow water habitats (i.e., up to 50 m depth), thermo-tolerance is influenced by local temperature fluctuations. investigations throughout a wider bathymetric range are, however, still lacking. this work aimed at testing if c. rubrum colonies belonging to different populations (shallow versus deep water populations) do show different sensitivity to thermal stress. to do so, corals dwelling at different depths (32 m and 100 m) were collected, kept in aquaria, and their responses to thermal stress was evaluated through two lab experiments. oxygen consumption at different temperature regimes (14°c, 21°c and 25°c) and coenosarc necrosis at 25°c (over a 9-day lasting experiment) were used as response variables. shallow water samples showed significant lower oxygen consumption and coenosarc necrosis compared to deep samples, at all temperatures, revealing a potential higher survival rate in stress conditions. significant differences in oxygen consumption between shallow and deep dwelling samples were detected in the 21°c treatment. significant differences in necrosis were observed after 5 days, emphasizing a good tolerance to temperature increase for deep colonies. implications of the obtained results for conservation of mediterranean red corals are illustrated and discussed. key words: red coral; thermal stress; oxygen consumption; animal forest. received: january 2018. accepted: june 2018. no nco mm er cia l u se on ly a. cau et al.14 the bathymetric range of distribution is comprised between 5 m and >1000 m depth (knittweis et al., 2016), with a distinction between shallow water populations (<50m depth) and deep-water population (>50m depth) (cau et al., 2016; costantini et al., 2010; santangelo and abbiati, 2001). such distinction is also reflected by the genetic differentiation that characterize shallow and deep dwelling populations. these differences are due to hidden barriers like depth-dependant environmental conditions and/or the thermocline, with a threshold in connectivity placed between -40 and -50 m of depth (cannas et al., 2016, 2015; costantini et al., 2011; costantini and abbiati, 2016). up to date, few studies had verified the occurrence of differential thermo-tolerance among colonies of c. rubrum dwelling in different locations within shallow populations’ bathymetric distribution range (i.e., comprised within 50 m depth) (previati et al., 2010; torrents et al., 2008). as deep dwelling populations are expected to experience lower thermal fluctuations, the question has been raised as to whether colonies belonging to the two different abovementioned environments could show different capabilities in the tolerance to induced thermal stress, estimated through oxygen consumption and coenosarc necrosis pattern. methods to test the null hypothesis that shallow and deep dwelling colonies of corallium rubrum do not have a different response to thermal stress, a total of 64 colonies were collected from central western coast of sardinia (cw mediterranean sea, italy). the two response variables evaluated to measure thermotolerance were coenosarc necrosis and oxygen consumption. sampling was performed by professional divers in spring 2011, at two different depths: 32 colonies were collected at 32 m depth (henceforth called the ‘shallow samples’) and 32 colonies at 100 m depth (henceforth called the ‘deep samples’). after collection, colonies were rapidly transported to the laboratory and stored in two 100liters aquaria (32 per each aquarium, 16 ‘shallow’ and 16 ‘deep’) at controlled temperature (14±0.5°c), salinity (37.5-38‰) and with constant water movement (1 cm s–1) generated by sterile water pumps. controlled conditions used during acclimation period were similar to those measured in situ. the acclimation period lasted a minimum of 14 days before the beginning of the experiments. a total of 48 colonies were randomly chosen from the two aquaria and used in the oxygen consumption experiments (24 shallow and 24 deep colonies), while the remaining 16 were used in the necrosis experiments. for the oxygen consumption experiment, the 42 colonies were put singularly inside glass chambers left open. the experimental design considered three treatments (fig. 1): 14°c (used as control), 21°c and 25°c. selected temperature treatments were intended to simulate temperature conditions recorded in the field for shallow populations and were used in the present experiments to identify the upper thermal limits of the two populations. fig. 1. experimental design showing three aquarium sets used to carry out the thermo-tolerance experiments performed in this study. the number of shallow and deep samples used in each experiment is reported in each aquarium; each colony from deep and shallow samples is stored in a separate chamber within each replicate. light and dark grey coloured bars represent the incoming and outgoing seawater flow. no nco mm er cia l u se on ly thermal stress of shallow and deep dwelling colonies of mediterranean red coral corallium rubrum (l., 1758) 15 the experimental conditions were obtained gradually, rising temperature throughout 48 h. when experimental conditions were reached, glass chambers were hermetically closed and the experiment started. for each treatment, four replicates were set up, each one composed by 4 colonies (2 shallow and 2 deep), stored separately in 4 glass chambers (fig. 1). oxygen consumption was estimated as mean oxygen consumption per hour during the treatment (duration 20 hours). oxygen level was measured with an oxymeter at the beginning and at the end of the experiment; overall, 96 measures of oxygen consumption (i.e., beginning/end of the experiment) were retrieved. colonies were used exclusively for a single replicate in a single treatment. in addition, in order to standardize as much as possible the number of polyps for colonies used into the oxygen consumption experiment, all colonies used belonged to the same size-class (i.e., 6-7 cm in height). for each replicate, an empty control chamber was included; the oxygen measures were performed also in the control chamber at the beginning and at the end of the experiment, in order to subtract recorded value from values registered in the chambers containing colonies. at the end of the treatment the dry mass of the coenosarcs (dmc) of each colony was measured: the colony was first weighted, then dried at 90°c for 24 h, and finally weighted again after the complete removal of coenosarc (that represents the living tissue that performs respiration). the difference between the two weight measures is the dry mass. the oxygen consumption of each colony was measured as mg o2 g dmc−1 h−1, which is a mean value per hour over the total time between the initial measurement and the final measurement. a two-way analysis of variance (anova, α=0.05) was performed to test the differences of the oxygen consumption between depths (shallow vs deep) and temperature regimes (14°c vs 21°c vs 25°c). prior to the analysis, cochran’s c test (α=0.05) was used to check the assumption of the homogeneity of variances and, whenever necessary, data were square rooted. where data transformation did not correct violations in the assumption of homogeneous variances, an alpha-level adjustment to 0.01 was used to compensate for increased type i errors (underwood, 1997). secondly, in order to focus on differences in consumption between depth samples (shallow vs deep) within every temperature regime, oxygen consumptions rates were compared with the t-test. both deep and shallow stressed colonies that were used for the oxygen consumption experiment were brought back at normal thermal conditions (14°c). survival rate was retrieved after 4 weeks, calculated as overall percentage of live colonies; colonies were considered alive even only if a small part of coenosarc showed living polyps. finally, in order to determine potential different survivor capability to long heat stress, 8 colonies (4 shallow and 4 deep) where kept at 25°c temperature regime during 9 days. every 24 h, two high definition pictures from fixed positions were taken for both sides of the colony. pictures were analysed through cpce software v.4.1 (kohler and gill, 2006) to determine the occurrence of necrosis on each colony. necrosis was expressed as percentage of denuded axis compared to the total surface of the colony, which was obtained by both photographed sides of every colony. eight remaining colonies not used for stress experiments (always left at 14±0.5°c) were used as control. results oxygen concentration at the beginning of the experiments varied from 7.49±0.13 mg o2 l–1 at 14°c to a minimum of 6.71±0.22 mg o2 l–1 at the highest temperature (25°c; fig. 2). for all controls, oxygen consumption did not change significantly at the beginning and at the end of each treatment (t-test, pvalue=ns). in the chambers containing colonies the consumption rates varied between temperatures (fig. 2); both shallow and deep colonies showed an increase in oxygen consumption at 21°c and, notwithstanding, a decrease at 25°c. focusing on the comparison between depths within treatments, the t-tests indicated that oxygen consumption was significantly higher in deep than in shallow colonies for the 21°c and 25°c regimes. the two-way anova (tab. 1) showed significant differences between depths and temperatures, fig. 2. oxygen consumption measured as mg o2 g dmc−1 h−1 (bars ± sd) for both deep (black bars) and shallow (grey bars) samples at 14°c, 21°c and 25°c. the t-test was performed to compare oxygen consumption between deep and shallow samples. ns, not significant; *p<0.05; **p<0.01; ***p<0.001. no nco mm er cia l u se on ly a. cau et al.16 highlighting also a significant positive interaction between the two factors. during the 9 days’ thermal stress experiment, differences in the necrosis pattern were observed on a daily base between deep and shallow colonies. at the end of the experiment the mean percentage of necrotic tissue was significantly higher (t-test, p<0.05) for the deep sample (84.41±11.51%) compared to the shallow sample (31.12±18.32%). observing the necrosis pattern progressing across days, the shallow sample showed a more linear trend with slow increment throughout all 9 days of experiment, while deep colonies increased the advancement of necrosis after day 5. in fact, comparing the percentage of necrotic pattern of the 2 samples for each day, significant differences are reported from day 6 and for all following days (figs. 3 and 4). as concerns the survival rate, 75% of shallow colonies survived thermal stress, while only 12.5% (1 colony over 8) of deep colonies survived. colonies from the control sample did not show any necrosis. fig. 3. mean occurrence of necrosis for sample, detected as percentage of denuded axis (points ± sd) for each day of treatment. test of significant differences between samples (black squares deep sample, grey circles shallow sample at same day). ns not significant; *p<0.05; **p<0.01; ***p<0.001. fig. 4. advancing necrosis in a shallow colony of c. rubrum under 25°c temperature regimes after 5 (left) and 9 days (right). the white area represents the necrotic tissue, measured with cpce software. tab. 1. results of the two-way anova analysis on oxygen consumption. analysis of variance source sum of squares df mean square f ratio p depth 0.0314 1 0.0314316 15.37 0.0003 temperature 0.7701 2 0.385084 188.32 0.0000 interactions depth x temperature 0.0387 2 0.0193 9.48 0.0004 residuals 0.0858 42 0.002 total (corrected) 0.9262 47 df, degree of freedom. no nco mm er cia l u se on ly thermal stress of shallow and deep dwelling colonies of mediterranean red coral corallium rubrum (l., 1758) 17 discussion several studies demonstrated that upper thermal limits of temperate corals are correlated with the environment and their previous stress history (ezzat et al., 2013; linares et al., 2013). the aim of this work was to test the existence of a differential tolerance to thermal stress between colonies of corallium rubrum belonging to different populations (shallow versus deep dwelling populations). previous studies (torrents et al., 2008; previati et al., 2010) investigated physiological response to thermal stress among colonies dwelling above and below the thermocline threshold, within shallow populations distribution range (up to 50m depth). our data dealt with a wider bathymetric range comprising both shallow (32 m depth) and deep dwelling colonies (100 m depth), providing empirical evidence of differential thermotolerance in c. rubrum samples from shallow and deep population. both oxygen consumption (fig. 2) and tissue necrosis showed significant differences between shallow and the deep samples (figs. 3 and 4). it is a remarkable finding that colonies (both deep and shallow) survived without showing a considerable increase in oxygen consumption and a marked necrosis until the temperature that was 11°c higher than the control temperature. shallow water samples showed a significant lower oxygen consumption and coenosarc necrosis at all temperatures compared to deep samples, revealing a potential higher survival rate under stress conditions. on the other hand, significant differences in oxygen consumption between shallow and deep dwelling samples were detected in the 21°c treatment, and after 5 days in the necrosis experiment, underlining, however, a good capacity of tolerance to temperature increase also in deep colonies. in particular, deep dwelling colonies of c. rubrum could be tolerant to water temperature increase that had already been observed in the deep waters of the mediterranean basin (haguenauer et al., 2013; torrents et al., 2008). observed differences in thermo-tolerance could be explained through the strong genetic differentiation that is known to occur between shallow and deep populations of c. rubrum. in this regard, haguenauer et al., (2013) investigated the expression of the heat shock protein 70, defining marginal populations living on the edge of the species distribution in the shallow environment (dwelling at ca. 5 m depth) as a fundamental genetic pool for evolution. future studies, considering that the different capability to respond to thermal stress of shallow red coral populations is genetically determined should extend the investigation to deeper populations. moreover, in the present study the highest differences in thermo-tolerance (between shallow and deep dwelling colonies) where found in the 21°c treatment, but they slightly decreased at the hottest temperature (25°c). similarly to our results, investigations conducted on other species showed different responses to thermal stress with depth. in eunicella singularis (esper, 1794), for example, colonies dwelling in the deeper range of the bathymetric distribution of the species (below 35m) showed a higher tolerance than shallower ones (ezzat et al., 2013; ferrierpagès et al., 2009); in addition, calcification and photosynthesis was compromised only after 14 days of high temperature exposure (ferrier-pagès et al., 2009). eunicella cavolinii (koch, 1887) showed a higher thermotolerance with no sign of necrosis after three days of exposition to 25°c (previati et al., 2010) but only an alteration of the coenenchyme colour. on the other side, paramuricea clavata (risso, 1826) proved to be highly sensitive to high temperatures, showing a peak of polyp activity at 18°c and nearly complete retraction at 25°c (previati et al., 2010; cerrano et al., 2000). in summary, the present study gives new insights suggesting the potential role of deep areas to act as ‘thermal refugia’ (places where colonies survive during periods of thermal stress; bongaerts et al., 2010); this aspect is of great importance for conservation of the species as climatic change is supposed further to warm the oceans and consequently affect increasingly shallow red coral populations in the very future (bramanti et al., 2013; cerrano et al., 2013). aknowledgements alessandro cau gratefully acknowledges university of sassari for the financial support of his post doctoral fellowship and the department of life and environmental science of university of cagliari for the logistic support for the aquaria experiments. in addition, all authors acknowledge all professional divers involved in the collection of samples. references bongaerts p, ridgway t, sampayo em, hoegh-guldberg o. 2010. assessing the “deep reef refugia” hypothesis: focus on caribbean reefs. coral reefs 29:1-19. bramanti l, movilla j, guron m, calvo e, gori a, dominguezcarrió c, grinyó j. lopez-sanz a, martinez-quintana a, pelejero c, ziveri p, rossi s, 2013. detrimental effects of ocean acidification on the economically important mediterranean red coral (corallium rubrum). glob. chang. biol. 19:1897-1908. cannas r, sacco f, cau a, coluccia e, follesa mc, cau a, 2015. new insights into connectivity patterns of mesophotic red coral (corallium rubrum) populations. hydrobiologia 759:63-73. cannas r, sacco f, cau a, cuccu d, follesa mc, cau a, 2016. genetic monitoring of deep-water exploited banks of the precious sardinia coral corallium rubrum (l., 1758): useful no nco mm er cia l u se on ly a. cau et al.18 data for a sustainable management. aquat. conserv. 26:236250. cau a, bramanti l, cannas r, follesa mc, angiolillo m, canese s, bo m, cuccu d, guizien k.,2016. habitat constraints and self-thinning shape mediterranean red coral deep population structure: implications for conservation practice. sci. rep. 6:23322. cau a, follesa mc, moccia d, alvito a, bo m, angiolillo m, canese s, paliaga em, orrù pe, sacco f, cannas r, 2015. deepwater corals biodiversity along roche du large ecosystems with different habitat complexity along the south sardinia continental margin (cw mediterranean sea). mar. biol. 162:1865-1878. cerrano c, bavestrello g, bianchi cn, cattaneo-vietti r, bava s, morganti c, morri c, picco p, sara s, schiaparelli s, siccardi a, sponga f, 2000. a catastrophic mass-mortality episode of gorgonians and other organisms in the ligurian sea (northwestern mediterranean), summer 1999. ecol. lett. 3:284-293. cerrano c, cardini u, bianchelli s, corinaldesi c, pusceddu a, danovaro r. 2013. red coral extinction risk enhanced by ocean acidification. sci. rep. 3:1457. costantini f, abbiati m, 2016. into the depth of population genetics: pattern of structuring in mesophotic red coral populations. coral reefs 35:39-52. 5. costantini f, rossi s, pintus e, cerrano c, gili jm, abbiati m, 2011. low connectivity and declining genetic variability along a depth gradient in corallium rubrum populations. coral reefs 30:991-1003. costantini f, taviani m, remia a, pintus e, schembri pj, abbiati m, 2010. deep-water corallium rubrum (l., 1758) from the mediterranean sea: preliminary genetic characterisation. mar. ecol. 31:261-269. crisci c, bensoussan n, romano jc, garrabou j, 2011. temperature anomalies and mortality events in marine communities: insights on factors behind differential mortality impacts in the nw mediterranean. plos one 6:e23814. di camillo, c.g., cerrano, c., 2015. mass mortality events in the nw adriatic sea: phase shift from slowto fast-growing organisms. plos one 10:e0126689. ezzat l, merle pl, furla p, buttler a, ferrier-pagès c, 2013. the response of the mediterranean gorgonian eunicella singularis to thermal stress is independent of its nutritional regime. plos one 8:e64370. ferrier-pagès c, tambutté e, zamoum t, segonds n, merle pl, bensoussan n, allemand d, garrabou j, tambutté s, 2009. physiological response of the symbiotic gorgonian eunicella singularis to a long-term temperature increase. j. exp. biol. 212:3007-3015. garrabou j, coma r, bensoussan n, bally m, chevaldonné p, cigliano m, diaz d, marschal c, pérez t, ribes m, romano jc, serrano e, teixido n, torrents o, zabala m, zuberer f, cerrano c, 2009. mass mortality in northwestern mediterranean rocky benthic communities: effects of the 2003 heat wave. glob. change biol. 15:1090-1103. garrabou j, perez t, sartoretto s, harmelin j, 2001. mass mortality event in red coral corallium rubrum populations in the provence region (france, nw mediterranean). mar. ecol. prog. ser. 217:263-272. haguenauer a, zuberer f, ledoux jb, aurelle d, 2013. adaptive abilities of the mediterranean red coral corallium rubrum in a heterogeneous and changing environment: from population to functional genetics. j. exp. mar. bio. ecol. 449:349-357. hughes tp, baird ah, bellwood dr, card m, connolly sr, folke c, grosberg r, hoegh-guldberg o, jackson jb, kleypas j, lough jm, marshall p, nyström m, palumbi sr, pandolfi jm, rosen b, roughgarden j, 2003. climate change, human impacts, and the resilience of coral reefs. science 301:929-933. knittweis l, aguilar r, alvarez h, borg ja, evans j, garcia s, schembri pj, 2016. new depth record of the precious red coral corallium rubrum for the mediterranean. proc. congr. int. commission for the scientific exploration of the mediterranean sea, kiel 41:456. kohler, k.e., gill, s.m., 2006. coral point count with excel extensions (cpce): a visual basic program for the determination of coral and substrate coverage using random point count methodology. comput. geosci. 32, 1259-1269. lejeusne c, chevaldonné p, pergent-martini c, boudouresque cf, pérez t, 2010. climate change effects on a miniature ocean: the highly diverse, highly impacted mediterranean sea. trends ecol. evol. 25:250-260. linares c, cebrian e, kipson s, garrabou j, 2013. does thermal history influence the tolerance of temperate gorgonians to future warming? mar. environ. res. 89:45-52. meinshausen m, smith sj, calvin k, daniel js, kainuma mlt, lamarque jf, matsumoto k, montzka sa, raper scb, riahi k, thomson a, velders gjm, van vuuren dp, 2011. the rcp greenhouse gas concentrations and their extensions from 1765 to 2300. clim. change 109:213-241. previati m, scinto a, cerrano c, osinga r, 2010. oxygen consumption in mediterranean octocorals under different temperatures. j. exp. mar. bio. ecol. 390:39-48. rossi s, 2013. the destruction of the ‘animal forests’ in the oceans: towards an over-simplification of the benthic ecosystems. ocean coast. manag. 84:77-85. rossi s, tsounis g, orejas c, padron t, gili jm, bramanti l, teixido n, gutt j, padrón t, 2008. survey of deep-dwelling red coral (corallium rubrum) populations at cap de creus (nw mediterranean). mar. biol. 154.533-545. santangelo g, abbiati m, 2001. red coral: conservation and management of an over-exploited mediterranean species. aquat. conserv. 259:253-259. torrents o, tambutté e, caminiti n, garrabou j, 2008. upper thermal thresholds of shallow vs. deep populations of the precious mediterranean red coral corallium rubrum (l.): assessing the potential effects of warming in the nw mediterranean. j. exp. mar. bio. ecol. 357:7-19. truesdale ga, downing al, lowden gf, 1955. the solubility of oxygen in pure water and sea-water. j. appl. chem. 5:53-62. underwood aj, 1997. experiments in ecology: their logical design and interpretation using analysis of variance. cambridge university press, cambridge. van vuuren dp, meinshausen m, plattner gk, joos f, strassmann km, smith sj, wigley tm, raper sc, riahi k, de la chesnaye f, den elzen mg, fujino j, jiang k, nakicenovic n, paltsev s, reilly jm. 2008. temperature increase of 21st century mitigation scenarios. p. natl. acad. sci. usa 105:15258-15262. no nco mm er cia l u se on ly layout 1 introduction ice cover in lakes has a significant impact on various air-water exchange processes. it reduces the sunlight penetration into water, which is necessary for photosynthesis; it also hinders heat transmission and water oxygenation. ice conditions on lakes determine the duration of navigation period as well as the possibility of transporting people and/or cargoes over the stable ice (karetnikov and naumenko, 2008; salo and nazarova, 2011; assel et al., 2004). the data on ice regime characteristics is applicable to climate models and can be used to predict the periods of freeze-up and break-up phases of lakes (salo and nazarova, 2011; baklagin, 2017). the formation and break-up of ice cover on large lakes depend greatly on the complex of meteorological processes, which occur over the water area of lakes (salo and nazarova, 2011; dibike et al., 2011); therefore, research related to the influence of climate factors on long-term variability of ice regime of lakes is of high interest. reduction of ice phenomena duration due to global warming observed over the past few decades is the main tendency in the long-term variability of ice regime of large lakes (brown and duguay, 2010; latifovic and pouliot, 2007; efremova et al., 2013; magnuson et al., 1990). according to livingstone (1997), the ice of large lakes serves as a sensitive indicator of climate change that is even more reliable than air temperature. with the water surface area of 9720 km2, lake onego is one of the biggest lakes in europe. identifying the patterns of ice regime formation of lake onego is of great importance for issues connected with organizing water transport between such settlements as petrozavodsk, medvezhyegorsk, kondopoga, povenets, vytegra, voznesenye, as well as on white sea-baltic canal and volga-baltic canal (salo and nazarova, 2011). long-term variability of ice regime of large lakes on the territory of the republic of karelia, including lake onego, has been described by efremova et al. (2013). the authors observed that during the years 1950-2009, the duration of ice phenomena on lake onego reduced by more than 20 days. identified variability patterns of air temperature influence on the formation and break-up of ice cover show that for lake onego, in particular, freeze-up dates are most accurately indicated by november-december average air temperature, and break-up dates by that of april-may. furthermore, a ±1°c fluctuation of average air temperature leads to a ±4-6 days shift of the freeze-up dates and a ±3-4 days shift of the break-up dates (efremova et al., 2013). it is important to note that these results were derived from visual ice cover data at the observation stations of the russian meteorological service and that they are relevant for small lakes, which can visually inspected by the observers. in case of lake onego, the observation is focused on the ice cover condition of petrozavodsk bay. its surface area is less than 2% of the total surface area of lake onego, which is not sufficient for estimating the indicative periods of the ice regime. long-term variability of the ice regime on lake onego was described by salo and nazarova (2011) based on air observations of ice for the period of 1955-1990 performed by air research office of the north-west, the federal service for hydrometeorology and environmental monitoring. based on the data, the correlation between indicative periods, the duration of ice phenomena, air temperature (weather station petrozavodsk), and nao (northern atlantic oscillation) index was traced. it is worth noting that the data obtained from ice surface mapping of lake onego used in the study (salo and nazarova, 2011) lacks sufficient time interval for proper evaluation of indicative dates of ice regime and the analysis of chronological ice coverage progress (510 air observations within the period of ice formation). article variations of indicative dates of ice regime on lake onego based on ground air temperature vyacheslav n. baklagin northern water problems institute, karelian research centre, russian academy of sciences, aleksander nevsky st. 50, 185030 petrozavodsk, republic of karelia, russia abstract the paper shows the changes in the dates (complete freeze-up, ±5 days/°c and complete ice clearance, ±3 days/°c) of the ice regime in lake onego depending on changes in average air temperature within the preceding two-month periods (autumn and spring). the regression equations for their calculation based on previous threeand four-month periods according to the 20002018 data are also provided. indicative dates of ice regime based on accumulated air temperatures within the ice period of lake onego were also established (early formation of ice phenomena, complete freeze-up phase, beginning of the break-up phase and complete ice clearance). together with the data on expected air temperature above the lake’s surface, these dependencies enable us to predict the indicative dates of the ice regime. no nco mm er cia l u se on ly variations of indicative dates of ice regime on lake onego based on ground air temperature 25 the maximum alternation rate of ice coverage of lake onego per 24 hours for the years 2000-2018 is 62.5% (registered by modis sensor, from 8th to 9th january, 2016). more reliable information on the ice cover of lakes is obtained from satellite observations (karetnikov and naumenko, 2008; assel et al., 2004; baklagin, 2018). for the last few years, daily satellite surveys of the earth using various ranges (visual, infrared, microwave), have collected large amounts of data including the data about snow and ice cover of the planet. for this reason, it is necessary to update previous studies (salo and nazarova, 2011; efremova et al., 2013) based on present satellite data and to confirm previously identified correlations. besides that, we need to analyze the potential influence of temperature conditions and some other meteorological factors (such as wind and snow cover depth) on ice cover formation and break-up on lake onego, since these issues have not been fully described in the available studies. the purpose of this study is to identify statistical correlations between meteorological parameters and the ice regime characteristics of lake onego calculated on the basis of satellite data within the years 2000-2018. methods identification of the ice regime in lake onego given the possible quick changes in areas of ice phenomena on lake onego (up to 63% per 24 hours), the ice regime characteristics of lake onego for the years 20002018 were calculated on the basis of daily time series of ice coverage obtained from satellite observations. in this study we used satellite data provided by national aeronautics and space administration, usa, nasa (device modis – the moderate resolution imaging spectroradiometer, with a spatial resolution of 250 m), national snow and ice data center, nsidc (4-6 km), center for satellite applications and research, noaa nesdis (46 km). the technique described by bakalgin (2018) was used to develop daily time series of ice cover of lake onego on the basis of these data series with minimizing faulty proportion while identifying ice cover (fig. 1). for the purposes of comprehensive assessment of the change of ice cover during the period of ice phenomena, sums of daily values of ice coverage for each period of ice phenomena were calculated using the formula: ∑ice=∑n(k=1)icek , where icek is ice coverage value (ice covfig. 1. dates of the beginning and the ending of phases and the duration (number of days) of ice freezing (1), complete freeze-up (2) and break-up (3) on lake onego for the period 2000-2018. no nco mm er cia l u se on ly v.n. baklagin26 erage is defined as the ratio of the ice cover area to the total lake area) in k-day of the period of ice phenomena, n –duration of the period of ice phenomena. values ∑ice and rici, i.e. the calculation technique that was used to estimate the ice regime of lake ladoga (karetnikov and naumenko, 2008) are similar, since values rici for each year were obtained by normalizing of value ∑ice to mean value ∑̅̅i̅c̅e̅ for multiyear period. however, in this study there is no need in normalizing values ∑ice for identification of statistical correlation. estimation of meteorological conditions above lake onego in this study we used daily data of mean air temperature, precipitation, and snow-cover depth for the years 2000-2017 from meteorological observation points near petrozavodsk and vytegra provided by all-russia research institute of hydrometeorological information world data centre (rihmi-wdc) (http://meteo.ru). data on daily mean air temperature and wind speed for the years 2000-2018 from meteorological observation points near petrozavodsk, medvezhyegorsk, vytegra, and pudozh were provided by national oceanic and atmospheric administration, usa, noaa (ncdc noaa) (ftp://ftp.ncdc.noaa.gov/pub/data/noaa/). meteorological conditions above lake onego were estimated by averaging of parameters measured at four observation points around the lake (fig. 2): near petrozavodsk (22820), medvezhyegorsk (22721), vytegra (22837), and pudozh (22831) (with index of the world meteorological organization, wmo). several observation points for estimating meteorological parameters were chosen to cover the variability of climatic conditions caused by the large dimensions of lake onego (from north to south, 248 km; from west to east, 96 km). for instance, the average air temperature according to rihmiwdc and ncdc noaa for the period of 2000-2017 at medvezhyegorsk is 3.11°с, in petrozavodsk -3.89°с, and in vytegra -4.06°с. the dates of the beginning and end of the periods of accumulated positive temperatures ∑t+ and negative temperatures ∑t– were derived from the condition: ∑tt 21 |tt |→max, where t1, t2 are the dates of beginning and end of the periods during the current hydrological year, tt is the average air temperature for the date t. method of determining the indicative dates of ice regime based on accumulated air temperatures this study assumes that accumulated negative air temperatures over the lake required for beginning ice phenomena formation ∑tfreezing and the complete freeze-up on the lake ∑tice, probably depend on thermal reserve of the lake before the beginning of the cold season in question. this thermal reserve of the lake depends on accumulated positive air temperatures ∑t+ of previous warm season. moreover, accumulated negative air temperatures over the lake required for the beginning break-up ∑tbreaking and the complete ice clearance on the lake ∑tfree, depend on accumulated negative air temperatures ∑t– of previous cold season. in the study, regression analysis was used to determine the dependencies of the accumulated air temperatures required for changing the phases of ice regime (indicative date) on accumulated air temperatures of previous period, i.e. the accumulated air temperatures required for changing the phases of ice regime (indicative date) is function from accumulated air temperatures of previous period ∑tind date (∑t). we can calculate the indicative date of the ice regime dateind based on a forecast of air temperatures for a period (for example, a week) and the calculated value ∑tind date (∑tfreezing, ∑tice, ∑tbreaking, ∑tfree). the indicative date of ice regime is determinate: fig. 2. the location of meteorological observation points on the coast of lake onego. no nco mm er cia l u se on ly variations of indicative dates of ice regime on lake onego based on ground air temperature 27 dateind = datecur + x (eq. 1) where datecur is the current date; x is the duration of the period from current date to the indicative date of the ice regime (days). value x can be determined based on the condition: ∑tind date ≈∑tcur date +∑xj =1 tj, (eq. 2) where ∑tcur date) is the accumulated air temperatures for period from the date of air temperature transition over 0°с to current date (°с); j is the counting number of days from current date; tj is the daily predicted air temperature of jth days (°с). the method of calculation resides in searching the value x where eq. 2 will be true. results for the analyzed range of years 2000-2018, the duration of ice phenomena period varied from 132 to 203 days with the average value of 171 days (from november to may), coefficient of variation is 10%, complete freeze-up period from 11 to 137 days, with the average value of 90 days, coefficient of variation is 36%. statistical characteristics of accumulated positive and negative air temperatures over lake onego are represented on fig. 3. for the years 2000-2018, the accumulated sums of positive air temperatures over the lake onego ∑t+max ranged from 1932°с to 2527°с, absolute values of negative air temperatures ∑t– max from 553°с to 1520°с. it is noteworthy that the average daily air temperature over lake onego in the years 2000-2018 was 0.9°с higher than in 1936-1999 (data provided by rihmi-wdc) (tab. 1). in addition, this increase in the average air temperature was mainly caused by frequent warm winters over the last years. over the years 20002018, the cold seasons differed from one another in terms of temperature conditions more considerably (variation coefficient of ∑t– max 32%) than the warm seasons (variation coefficient of ∑t+max 7%). according to the calculations for the years 2000-2018, the dates of air temperature transition over 0°с to negative values over lake onego were within the range from 14th october to 25th november, and to positive values from 4th march to 21st april. the average period with positive air temperatures (warm season) lasted 220 days, and with negative air temperatures (cold season) 144 days. correlation analysis showed, that the following values: ∑ice, the duration of freeze-up d (days) and the total period of ice phenomena l (days) on the lake onego have close correlation (l to a lesser extent) with annual average air temperature over the lake t̅ (paired correlation coeffifig. 3. integral curves of positive (a) and negative (b) air temperatures over lake onego with the range of sample for the period 2000-2018. tab. 1. temperature conditions over lake onego at different time periods. period average air temperature (°с) difference (°с) 1936-1999 2000-2018 year 2.7 3.6 0.9 warm season 9.6 10.4 0.8 cold season –8.2 –6.8 1.4 no nco mm er cia l u se on ly v.n. baklagin28 cients -0.89, -0.83 and -0.50, respectively). the regression analysis revealed polynomial relationships of ∑ice and d with annual average air temperature over the lake t̅ : ∑ice= –0,548 ∙ t̅ 2–17,566 ∙ t̅ +188,649 (r2=0.8; p<0.05) (eq. 3) d= –3,894 ∙ t̅ 2–3,782 ∙ t̅ +157,531 (r2=0.7; p<0.05) (eq. 4) updating air temperature – ice regime relationships the starting date of ice phenomena formation on lake onego was calculated by multiple regression from monthly average air temperatures: dfreezing= 0,249 ∙ t̅xii2 – 0,061 ∙ t̅xi2 – 0,444 ∙ t̅x2 + 4,253 ∙ t̅ix2 + 3,579 ∙ t̅xii + 1,974 ∙ t̅xi + 4,122 ∙ t̅x – 79,643 ∙ t̅ix + 422,458 (r2=0.81; p<0.05) (eq. 5) where dfreezing is the duration of the period from 1st october to the beginning of ice phenomena formation, (days); t̅i is the average air temperature of ith month over the lake in °с. strong correlation (r= 0.76; p<0.05) was also identified between the average air temperature for a two-month period from november to december t̅xi-xii and the dates of ice complete freeze-up on the lake for 2000-2018. the dependency of changing of these dates from the value t̅xi-xii was obtained (fig. 4). the slope factors of dependences are coherent with the results of a similar study on the ice regime of lake onego for the period of 1950-2009 (efremova et al., 2013) (tab. 2). the average air temperature over the lake from december to january (t̅xii-i) had the strongest correlation (r=0.88; p<0.05) with the ice complete freeze-up on lake onego for the period of 2000-2018. (tab. 2). however, values t̅xi-xii were used to define relationships in the study by efremova et al. (2013). considering air temperatures of a 4-month period (october-january) instead of a 3-month period (october-defig. 4. dependences of indicative dates of ice regime on lake onego on air temperature for the years 2000-2018, where a) dates of complete freeze-up depending on the average air temperature for period from november to december; b) dates of complete freeze-up depending on the average air temperature for period from december to january; c) dates of complete ice clearance depending on the average air temperature for period from april to may; d) dates of complete ice clearance depending on the average air temperature for april. 1, actual values; 2 linearly regression function. no nco mm er cia l u se on ly variations of indicative dates of ice regime on lake onego based on ground air temperature 29 cember) as used by salo and nazarova (2011) for estimating the complete freeze-up on lake onego for 1955-1990, as well as applying the polynomial multiple regression helped to improve the explanatory power of the regression model (r2 rose from 0.74 to 0.89): dice=0,380 ∙ t̅i2+ 0,073 ∙ t̅xii2 + 0,267 ∙ t̅xi2 + 0,692 ∙ t̅x2 + 10,295 ∙ t̅ i + 3,466 ∙ t̅xii + 2,109 ∙ t̅xi – 3,881 ∙ t̅x + 187,491 (r2=0.89; p<0.05) (eq. 6) where dice is the duration of the period from 1st october to the beginning of complete freeze-up phase, days. the beginning of ice cover break-up could be determined by air temperature from january to march: dbreak = –0,089 ∙ t̅iii2 – 0,385 ∙ t̅ii2 – 0,130 ∙ t̅i2 – 2,344 ∙ t̅iii – 8,701 ∙ t̅ii – 4,412 ∙ t̅i – 27,842 (r2=0.63; p<0.05) (eq. 7) where dbreak is the duration of the period, starting with 1st of march, until the beginning of ice cover break-up, days. a lower negative correlation (r=-0.52; p<0.05) was found between the dates of the complete ice clearance on lake onego and the april-may average air temperature t̅iv-v. considering april air temperatures only slightly strengthened the correlation (tab. 2). the best multiple regression model for estimating the dates of complete ice clearance included air temperatures from march to may: dfree = 0,328 ∙ t̅v2 – 0,096 ∙ t̅iv2 – 0,126 ∙ t̅iii2 – 6,910 ∙ t̅v – 2,925∙ t̅iv – 2,370 ∙ t̅ iii + 79,620, (r2=0.71; p<0.05) (eq. 8) where dfree is the duration (in days) of the period from 1st april to complete ice clearance. averages of the absolute deviations of the calculated values from the actual values were as follows: determining the date of the beginning of the ice formation: 3-4 days; determining the date of freeze-up phase: 4-5 days; determining the date of the beginning break-up phase: 23 days; determining the date of complete ice clearance on the lake: 7-8 days (fig. 5). the effect of the accumulated positive and negative air temperatures on ice regime the values ∑tice, ∑tfree for the years 2000-2018 showed a wide variation (tab. 3). no correlation was detected between ∑tice and ∑t+max (r = –0.01). ∑tice and ∑tfreezing had strongest correlations (correspondingly, r = 0.70; p<0.05 and r = -0.57; p<0.05) with the accumulated positive air temperatures for the last 55 days of the warm season (before the transition to negative values), ∑t55. on the basis of regression analysis, equations were obtained connecting ∑tfreezing and ∑t55, and ∑tice and ∑t55 for lake onego: ∑tfreezing = 7,195 ∙ 10–5 ∙ ∑t552 – 0,092 ∙ ∑t55 – 7,768. (r2=0.33; p<0.05) (eq. 9) ∑tice = 1,1311 ∙ 10–3 ∙ ∑t552 – 1,218 ∙ ∑t55 – 116,855. (r2=0.53; p<0.05) (eq. 10) the analysis connecting ∑t– max and ∑tfree , ∑t– max and ∑t–break showed a close statistical correlation (by the cheddok scale), with pair correlation coefficient between ∑t– max and ∑tfree and ∑t– max and ∑t–break equal to -0,75 and -0,78, respectively. regression analysis connecting ∑t– max and ∑t–break , and ∑t– max and ∑tfree showed the following results: ∑t–break = – 4,8389 ∙ 10–7 ∑t– max3 – 0,0016 ∙ ∑t– max2 – 1,6623 ∙ ∑t– max – 527,852 (r2=0.72; p <0.05) (eq. 11) ∑tfree = – 1,9775 ∙ 10–7 ∙∑t– max3 – 0,0008 ∙ ∑t– max2 – 1,1202 ∙ ∑t– max – 268,835 (r2=0.66; p<0.05) (eq. 12) the mean values of the absolute deviations of accumulated air temperatures from the actual values were as tab. 2. changes in indicative dates of the ice regime (complete freeze-up and complete ice clearance dates) on lake onego per ± 1°c change in average air temperature in different periods. period pair correlation coefficient date changes per ± 1°c change in mean air temperature, days 1950-2009 2000-2018 november-december 0.76 ±4-6 ±5 december-january 0.88 ±5 april-may -0.52 ±3-4 ±3 april -0.59 ±3 no nco mm er cia l u se on ly v.n. baklagin30 follows: 5-6°c at the beginning of the period of formation of ice phenomena; –38-39°c at the time of complete freeze-up; 12-13°c at the beginning of the period of ice break-up; –27-28°с at the moment of complete ice clearance of the lake from ice. discussion calculation results of average durations of warm and cold seasons corresponds to the results of the research (nazarova, 2013). it is important to note that the average duration of ice phenomena period of lake onego (171 days) is longer than the period of negative air temperatures over the lake. prediction of indicative dates of the ice regime of lake onego using the regression equations (5)-(8) for calculations dfreezing, dice, dbreak, and dfree is not justified because, in most cases, the predicted date precedes the periods used as input data for which average air temperatures are calculated. for example, the average statistical date for the complete freeze-up phase on lake onego over the period 2000-2018 is january 16th, but in some years the complete fig. 5. actual (1) and predicted on the basis of the provided regression models (2) indicative dates of the beginning of the ice phenomena formation (a); complete freeze-up phase (b); beginning of ice cover break-up (c); and complete clearing of ice (d) on lake onego for the period 2000-2018. tab. 3. accumulated air temperatures over the water area of lake onega within the indicative dates of the ice regime for different time periods. data for 1955-1990 are from the study by salo and nazarova (2011). values periods 1955-1990 2000-2018 accumulated air temperatures over the waters of lake onego at the beginning of the formation of ice cover ∑tfreezing (°с) –43 –15 at the complete freeze-up phase ∑tice (°с) –490 –290 –500 –275 at the beginning of break-up phase ∑tbreak (°c) 0-115 at the moment of complete ice clearance on the lake ∑tfree (°c) 200-310 87-294 no nco mm er cia l u se on ly variations of indicative dates of ice regime on lake onego based on ground air temperature 31 freeze-up phase occurred in mid-late december (fig. 1), while the equation for calculating the dice value includes the average value of january air temperatures, which obviously do not affect the dice value. the strong correlation (0.88) in this case can probably be explained by the reverse effect of the ice cover on the air temperature over its surface. the calculated intervals of the values ∑tice , ∑tfree for the period of 2000-2018 were consistent with the results by salo and nazarova (2011) (tab. 3). probably strong correlations between ∑tice with ∑t55 and ∑tfreezing with ∑t55 is caused by the fact that high positive air temperatures over lake onego in autumn (on the average during the period 2000-2018 september-october) inhibit water cooling in the lake, keeping the heat accumulated during the summer. therefore, the bigger value of accumulated negative air temperatures ∑t– is required for further drop of the water surface temperature to 0°с and for ice formation. the described phenomenon can occur in autumn in case of a sudden winter which is followed by transition from moderate positive temperatures (5-10°с) to negative ones, omitting small positive temperatures (0-5°с). the obtained equations for calculating ∑tfreezing, ∑tice, ∑t–break, ∑tfree taking into account data on expected air temperature, can potentially be used to predict the indicative dates of the ice regime of lake onego. it should also be noted that this study notes a shift in the period affecting the date of formation of ice cover of lake onego for the years 2000-2018, a month ahead comparing with the periods in question (efremova et al., 2013; salo and nazarova, 2011) (three months and two months, respectively). perhaps this phenomenon is associated with the late dates of the freeze-up phase on lake onego in 2000-2018 (on average january 16th) than in previous years, as a result of this tendency towards a reduction in duration of the period of ice phenomena on large lakes due to global warming (brown and duguay, 2010; latifovic and pouliot, 2007; efremova et al., 2013, magnuson et al., 1990). in addition, as already noted, calculations of the characteristics of the ice regime of lake onego in this paper (efremova et al., 2013) were made on the basis of observations of the condition of the ice cover of the petrozavodsk bay, which due to its morphological structure is covered with ice much earlier than the water area of the lake as a whole. moreover, significant differences in the values of the lower limit of the interval of value ∑tfree (tab. 3) is due to the presence of abnormally low ∑t+ values in 2013 and 2014, when a complete ice clearance of lake onego was recorded (∑tfree, respectively, 87°c and 92°c). in these years, abnormally warm winters preceded the break-up phase the accumulated negative air temperatures during the cold season ∑t– max had minimum values for the years 2000-2018 (about 555°c), with an average value of -970°c over this period. conclusions the dependences of the indicative dates of the ice regime of lake onego on air temperature are generally consistent with the results obtained earlier in (efremova et al., 2013; salo and nazarova, 2011). however, in 20002018 there was one-month ahead shift in the period affecting the date of the formation of ice cover in comparison with the second half of the 20th century, considered in the studies by efremova et al. (2013) and salo and nazarova, 2011). this indicates climate change in recent decades, which contributes to the late winter onset and, as a consequence the shift in the freeze-up dates, which is consistent with the concept of global warming. therefore, it can be concluded that the models of the formation of the ice cover of lake onego presented in papers (efremova et al., 2013; salo and nazarova, 2011) require some adjustment to be applicable nowadays. the equations for calculating the indicative dates of the ice regime (5)-(8) are hardly applicable for forecasting, but can be used for diagnostic purposes, for example, to re-design the long-term time series of the characteristics of the ice regime of lake onega based on the available daily data on air temperature. this is particularly true for the first half of the 20th century, when there are only fragmentary data on the condition of the ice cover of the petrozavodsk bay of lake onego and rare air observations due to the lack of appropriate technical means (satellite observations). the indicative dates of the ice regime of lake onego are potentially predictable on the basis of the equations derived in the paper. this may be of practical use in planning the navigation period and the organization of waterways. acknowledgments the study was fulfilled with the financial support of the grant of the president of the russian federation mk3379.2018.5. references assel r, drobot s, croley ii te, 2004. improving 30-day great lakes ice cover outlooks. j. hydrometeorol. 5:713-717. baklagin vn, 2017. selection of parameters and architecture of multilayer perceptrons for predicting ice coverage of lakes. ekol. bratislava 36:226-234. baklagin vn, 2018. variability of the lake onega ice coverage in the period 2000-2018 according to the satellite data. ice snow 58:552-558. brown lc, duguay cr, 2010. the response and role of ice cover in lake-climate interactions. prog. phys. geog. earth environ. 34:671-704. dibike y, prowse t, saloranta t, ahmed r, 2011. response of no nco mm er cia l u se on ly v.n. baklagin32 northern hemisphere lake-ice cover and lake-water thermal structure patterns to a changing climate. hydrol. process. 25:2942-2953. efremova tv, palshin ne, zdorovennov re, 2013. long-term characteristics of ice phenology in karelian lakes. est. j. earth sci. 62:33-41. karetnikov sg, naumenko ma, 2008. recent trends in lake ladoga ice cover. hydrobiology 599:41-48. latifovic r, pouliot d, 2007. analysis of climate change impacts on lake ice phenology in canada using the historical satellite data record. remote sens. environ. 106:492-507. livingstone dm, 1997. break-up of alpine lakes as proxy data for local and regional mean surface air temperatures. climatic change 37:407-439. magnuson jj, benson bj, kratz tk., 1990. temporal coherence in the limnology of a suite of lakes in wisconsin, u.s.a. freshwater biol. 23:145-159. nazarova le, 2013. climate change and fluctuations in the karelian-kola region. environment. technology. resources. proceedings 9th int. scientific practical conf. 1:53-56. salo ya, nazarova le, 2011. [multiannual variability of the onega lake ice regime in conditions of variability of the regional climate].[article in russian]. processes russian geogr. soc. 143:50-55. corresponding author: slava.baklagin@mail.ru key words: ice regime; lake onego; air temperature. received: 2 april 2019. accepted: 14 june 2019. this work is licensed under a creative commons attribution noncommercial 4.0 license (cc by-nc 4.0). ©copyright: the author(s), 2019 licensee pagepress, italy advances in oceanography and limnology, 2019; 10:8198 doi: 10.4081/aiol.2019.8198 no nco mm er cia l u se on ly layout 1 introduction habitat complexity is defined as the heterogeneity in the arrangement of physical structure in the habitat surveyed (sensu lassau and hochuli, 2004) and it represents one among the most important ecological factor in shaping structure and community dynamics. among others, it influences fish abundance, diversity in terms of species richness and composition (jones, 1988; bell and galzin, 1984; roberts and ormond, 1987; bell et al., 1991; hixon and beets, 1993; warfe and barmuta, 2004; harvey et al., 2005; willis et al., 2005; mangano et al., 2017). a particular relationship has been reported for several natural environments between the habitat complexity and animal community structure or assemblage compositions (i.e. both numbers of individuals and numbers of species; luckhurst and luckhurst, 1978; roberts and ormond, 1987; mcclanahan, 1994; mccormick, 1994; öhman and rajasuriya, 1998; gratwicke and speight, 2005; garcia charton and pérez ruzafa, 2008; porporato et al., 2014; mangano et al., 2015). the main mechanism invoked to explain it, is a reduction of predation pressure due to the increased amount of refuge available to prey species (hixon and beets, 1993; macpherson, 1994; caley and st. john, 1996; almany, 2004a). increase in available refuges due to enhanced substrate topography also has been shown to reduce competition for space (hixon and menge, 1991; almany, 2004b) as well as adding to niche dimensionality (macarthur and levins, 1967), both of which potentially increase fish abundance and distribution. the same pattern among spatial complexity, fish abundance and species richness has also been reported for artificial habitats such as, for instance, extractive platforms (chang et al., 1977; higo et al., 1980; buckley, 1982; shulman, 1984; chandler et al., 1985; roberts and ormond, 1987; gorham and alevizon, 1989; hixon and beets, 1989; bohnsack et al., 1991; love and york, 2006). surprisingly, the largest amount of these evidence has been collected outside the mediterranean sea, where in spite of the large number of oil and gas extractive platforms, this aspect is still poorly studied (fabi et al., 2002, 2004; consoli et al., 2007, 2013; andaloro et al., 2011, 2012; scarcella et al., 2011; mangano and sarà, 2017). the extraction of fossil fuels from offshore fields has strongly increased over the last decades to meet the global growing demand for energy (ghisel, 1997; terlizzi et al., 2008), this implies that the number of offshore platforms has increased the world over and, most probably, it will further increase in the future (de luca, 1999; pulsipher and daniel, 2000). advances in oceanography and limnology, 2018; 9(2): 59-67 article doi: 10.4081/aiol.2018.7918 this work is licensed under a creative commons attribution-noncommercial 4.0 international license (cc by-nc 4.0). the influence of habitat complexity on fish assemblages associated with extractive platforms in the central mediterranean sea pierpaolo consoli,1 maria cristina mangano,2* gianluca sarà,2 teresa romeo,1,3 franco andaloro1 1stazione zoologica anton dohrn, centro interdipartimentale della sicilia, milazzo (me); 2department of earth and marine sciences, university of palermo; 3institute for environmental protection and research (ispra), bio-cit, palermo, italy *corresponding author: mariacristina.mangano@gmail.com abstract in this work the influence of habitat complexity on fish assemblages associated with extractive platforms in the mediterranean sea was investigated. more specifically, at large spatial scale we tested the differences in fish assemblage between 4-legs vs 8-legs platforms, whereas at medium scale we evaluated, within each platform, the differences between internal structures with increasing complexity degrees (respectively: the water volume without any pillar complexity “0”; the junction of two pillars “1”; the junction of four pillars “2”). both univariate and multivariate analyses showed highly significant differences for each of the tested factors, as well as for their interaction. in general, at both medium and large spatial scales, mean species richness and abundance were positively correlated with the increasing habitat complexity with the highest values associated with 8-legs platforms and with the most complex internal structures within each platform. according to our findings, a more complex structure is able to attract more fish species and specimens than a less complex one, supporting previous studies carried out on different man-made structures outside the mediterranean sea. the study will integrate the still poor available knowledge baseline on the attractive potential of extractive platforms with strong implications for the environmental management under the incoming light of decommission in the basin. key words: artificial habitat; underwater visual census; gas platform; species richness. received: november 2018. accepted: november 2018. no nco mm er cia l u se on ly p. consoli et al.60 then, understanding the role played by offshore platforms in shaping multi-level marine ecosystem’s dynamics is becoming pressing as offshore platforms are acquiring increasing importance worldwide for its implications on marine biodiversity (mangano and sarà, 2017). the aim of the present study was to evaluate the influence of habitat complexity on fish assemblages associated with extractive platforms in the ionian sea (mediterranean sea). the obtained outcomes integrate the still poor available knowledge baseline on the attractive potential of these human-made structures with interesting rebounds for the environmental management in a context foreseen of decommission in the basin. in doing so, we tested whether different complexity degrees affected the associated fish assemblages across two different spatial scales. accordingly, 1) we tested the difference in fish assemblage at large scale (̴ 10 km) between two different levels of complexity (4-legs vs 8-legs platforms) and 2), at medium scale (̴ 100 m) testing the difference in fish assemblage between internal structures, of such platforms, with different spatial complexity. methods study sites the study was carried out during one week in may 2006 at three offshore gas platforms (luna a, luna b and hera lacinia) located in the southern ionian sea (central mediterranean sea) respectively, 5.3, 6.2 and 2.6 km offshore (fig. 1). two of them (luna a and luna b) were 8-leg platforms while the third one (h. lacinia) was a 4leg platform. all these platforms lie on a sandy seabed and are fixed to the sea floor by concrete or steel legs, which are connected by an assemblage of cross beams. the platforms were colonized by several foulers that generally provide crevices, refuges and food to cryptic and nekto-benthic fish species. the most abundant sessile species was the bivalve mytilus galloprovincialis followed by balanids, ostrea sp., and arbacia lixula (p. consoli, personal observation). habitat complexity for each of the three platforms, internal structures with increasing complexity degrees (hereafter complexity 0, 1 and 2; fig. 2), were identified and corresponded to: 0=the water volume without any pillar; 1=the junction of two pillars; and 2=the junction of four pillars, respectively. fish species and their abundances were recorded by underwater visual censuses (uvc) by deploying the “mobile point count” (mpc) technique performed at a depth between 0 and 12 meters. this technique, specifically designed for offshore platforms by rilov and benayahu (2000) and applied by consoli et al. (2007, 2013) and andaloro et al. (2011, 2012) in the mediterranean sea, was chosen as it is highly reliable in studying species strictly associated with the pillars and to detect benthic and cryptic species (andaloro et al., 2011, 2012; consoli et al., 2007, 2013). the diver, turning around each unit and looking at towards the pillar, counted all fishes occurring up to 3 m from the pillar. first, the diver recorded the more conspicuous and easily identifiable fishes from a maximum distance of 3 meters from the pillar (so that to have an entire view of the census unit) and then straight after, approached to the pillar, and counted the benthic and crypto-benthic species. the total censused volume for complexity 0 corresponded to a cylinder of 7 m of diameter and 6 m height (~231 m3). as regard complexity 1 and 2, the censused water volume was obtained subtracting the volume of the pillars (1 m of diameter) from 231 m3. the resulting volumes for complexities 1 and 2 were 224 and 219 m3, respectively. as a main consequence, data of abundance were standardized to the maximum censused volume (231 m3) in order to compare censuses performed next to the different spatial complexity structures. fortyfig. 1. study area located in the ionian sea off crotone (calabria, italy). no nco mm er cia l u se on ly the influence of habitat complexity on fish assemblages associated with extractive platforms in the central mediterranean sea 61 eight censuses were performed for each level of medium scale complexity at each platform, leading to a total of 432 observations in the data set. statistical analyses the sampling design included 2 factors: i) large scale complexity (lsc) was a fixed factor in the analysis with 2 levels of large-scale complexities (as expressed by: 4-leg and 8-leg platforms). ii) medium scale complexity (msc) was a fixed factor in the analysis with 3 levels: complexity 0, 1 and 2 according to the rationale presented before and represented in fig. 2. on this basis, a two-ways permutational analysis of variance (permanova; anderson, 2001, mcardle and anderson, 2001) was performed on abundance data to test the null hypothesis of no significant differences between fish assemblages associated with increasing habitat complexities, at two different spatial scales. the analysis was based on bray-curtis dissimilarities, calculated on log-transformed fish assemblage matrix. each term of the analysis was tested using 9999 random permutations of appropriate units (anderson and ter braak, 2003). significant terms that were relevant to our hypothesis were investigated using a posteriori pair-wise comparison with the permanova t-statistic and 9999 permutations. furthermore, we tested the effect of response variables on community metrics and in doing so, we modelled overall fish abundance and species richness through a permutational univariate analyses of variance (permanova; anderson, 2001; mcardle and anderson, 2001). here we used the euclidean distance instead the bray-curtis similarity index. thus, the same f-statistics were calculated, but p-values were obtained by permutation. finally, the simper similarity percentage procedure (clarke, 1993) was used to identify the fish species that most contributed to the differences among spatial complexities at medium and large spatial scale. all the analyses were performed using primer 6 software package with permanova+add-on (anderson et al., 2008). results in tab. 1 mean abundances and standard errors of each species are showed for lsc and msc factors. overall 15 fish taxa belonging to 6 families were recorded in the study area. most of the recorded species were nektobenthonic, while only 5 pelagic species were observed. in term of species richness, sparids were the most important family being represented by six species whereas the most abundant species were boops boops, anthias anthias and chromis chromis. permanova of the total fish assemblage (abundance data) showed highly significant differences for each factor considered in the analysis (tab. 2) and also for the interaction between factors lsc and msc (permanova, p=0.001). furthermore, pairwise comparisons showed that significant differences occurred between fish assemblages in every msc comparisons within each lsc level (p<0.001). the greatest t-values were observed between msc 0-level and 2-level at both 4and 8-legs platforms (t=3.85 and t=5.26, respectively; tab. 2). permanova on overall abundance and species richness mirrored the results of multivariate analysis (tab. 2). according to the large-scale complexity, the highest values of both metrics were associated with 8-legs platforms (s=2.62 and 1.58, n=237 and 56, respectively at 8-legs and 4-legs platforms; fig. 3). fig. 2. internal structures with increasing degrees of complexity; identified and corresponded to: 0=the water volume without any pillar; 1=the junction of two pillars; and 2=the junction of four pillars, respectively. no nco mm er cia l u se on ly p. consoli et al.62 as regards mean species richness, significant differences were found, at each platform, among msc levels (tab. 2) and the highest values were always associated with structures of compl 2 (fig. 3; tab. 2). looking at the t value, the greatest differences occurred between complexity 0 2 (t=17.747 and t=9.7555, at 8legs and 4-legs platforms, respectively, tab. 2). a similar pattern was observed, at 8-legs platforms, also for the mean abundance, whereas, at the 4-legs platform, the highest value was associated with level 0 of msc (compl 0; fig. 3). simper procedure pinpointed some fish taxa as the major contributors to the dissimilarities among spatial complexities. high densities of boops boops, anthias anthias and chromis chromis characterized the censuses carried out nearby the most complex structures, both at large and medium spatial scale (tab. 3). discussion fish assemblages associated with increasing habitat complexities showed differences in terms of species richness, abundance and assemblages structure. this tab. 1. mean species abundances and standard errors (± se) per sample unit (230.91 m3) for each level (compl 0, 1 and 2) of complexity factors at 4-legs vs 8-legs platforms. platform ecological lsc 4-legs 8-legs complexity category msc compl 0 compl 1 compl 2 compl 0 compl 1 compl 2 mean se mean se mean se mean se mean se mean se anthias anthias nb 31.53 11.85 86.14 14.39 187.00 21.95 boops boops p 30.00 6.99 24.54 8.88 23.84 6.29 31.31 7.04 56.85 11.60 137.29 19.23 chromis chromis nb 0.42 0.42 1.83 0.62 18.90 5.21 5.81 1.78 18.77 3.05 57.75 7.92 diplodus sargus nb 0.07 0.04 0.01 0.01 0.01 0.01 diplodus vulgaris nb 0.02 0.02 0.51 0.16 0.94 0.38 0.82 0.31 1.01 0.22 oblada melanura p 0.94 0.59 0.09 0.09 0.81 0.46 13.69 4.10 0.43 0.43 0.92 0.83 sarpa salpa nb 0.04 0.04 0.01 0.01 seriola dumerili p 0.02 0.02 serranus cabrilla nb 0.02 0.02 0.11 0.06 0.01 0.01 serranus scriba nb 0.02 0.02 spicara flexuosa p 26.15 9.12 13.00 5.17 14.06 5.92 4.24 1.39 13.58 5.48 32.69 7.91 spondyliosoma cantharus nb 0.02 0.01 0.02 0.02 0.10 0.04 thalassoma pavo nb 1.35 0.48 1.74 0.46 0.01 0.01 2.74 0.59 10.98 1.42 trachurus spp. p 9.38 4.25 0.86 0.60 0.44 0.44 1.77 1.12 1.92 1.15 13.44 6.66 lsc, large scale complexity; msc, medium scale complexity; nb, necto-benthonic; p, pelagic. tab. 2. results of permanova tests analysing the effect of lsc and msc factors on fish assemblage (multivariate test), species richness and fish abundance (univariate tests). results of pair-wise tests performed for the interaction factor “lsc x msc” are also reported. source fish assemblage species richness abundance df ms f p ms f p ms f p lsc 1 75,363 4.0051 0.002 102.78 50.627 0.001 3.14e+06 54.959 0.001 msc 2 52,279 16.826 0.001 133.95 113.99 0.001 2.15e+06 25.179 0.001 lsc x msc 2 18,817 6.0562 0.001 10.616 9.0341 0.001 2.14e+06 25.024 0.001 res 426 3107 1.1751 1.82e+07 8-legs platforms t p t p t p compl 0. compl 1 3.8476 0.001 10.934 0.001 3.9496 0.001 compl 0. compl 2 5.2578 0.001 17.747 0.001 9.1615 0.001 compl 1. compl 2 2.754 0.001 5.2947 0.001 6.3641 0.001 4-legs platforms compl 0. compl 1 2.3831 0.001 2.8711 0.001 1.2443 0.226 compl 0. compl 2 3.8503 0.001 9.7555 0.001 0.31125 0.772 compl 1. compl 2 1.8512 0.009 3.9344 0.001 0.95222 0.349 df, degree of freedom; lsc, large scale complexity; msc, medium scale complexity. no nco mm er cia l u se on ly the influence of habitat complexity on fish assemblages associated with extractive platforms in the central mediterranean sea 63 result was observed at both investigated spatial scales. in particular, as far as medium spatial scale is concerned, a positive relationship was observed between increasing habitat complexity and mean species richness at both levels of large spatial complexity (4and 8-legs platforms). the same pattern was detected for mean fish abundance at the most complex platforms, while at 4-legs platform, a clear pattern was not observed since the highest values were not associated with the most complex internal structures. in this less complex platform, internal structure, corresponding to different degree of medium spatial scale complexities, are usually closer to each other compared with those at 8-legs platforms. then, fishes probably, could not be able to distinguish these different degrees of spatial complexities. mean fish abundance and species richness resulted positively correlated with increasing complexities also at large spatial scale. these results strengthen and confirm observations made in previous studies carried out on different man-made structures such as artificial reefs (roberts and ormond ,1987; hixon and beets, 1989; chang et al., 1977; higo et al., 1980; buckley, 1982; gorham and alevizon, 1989; bohnsack et al., 1991; charbonnel et al., 2002, gratwicke and speight, 2005), fringing reef (roberts and ormond 1987), shipwrecks (chandler et al., 1985; fagundes-netto et al., 2011; consoli et al., 2015) and extractive platforms (love et al., 2003, 2010, 2012; love and york, 2006; rilov and benayahu, 1998, 2000, 2002; rooker et al., 1997; consoli et al., 2013). all these studies proved a positive relationship between fish species-richness/abundance and the increasing habitat complexity. after all, it is well known that these artificial habitats promote the aggregation of fishes that would otherwise be dispersed across larger areas of the ocean, a result of peculiar interest in the mediterranean basin, locally characterized by a very peculiar hydrodynamic system (hastings et al., 1976; aabel et al., 1977; driessen, 1985; gallaway et al., 1981; bohnsack and sutherland, 1985; love and westphal, 1990; bull and kendall, 1994; kasprzak, 1998; minton and heath, 1998; jørgensen et al., 2002; løkkeborg et al., 2002; love et al., 2003; love and york, 2006; andaloro et al., 2011, 2012; consoli et al., 2007, 2013; capodici et al., 2018). in particular, as regard extractive platforms, as these structures extend throughout the entire water column, their effects are not confined to demersal fishes, but also involve pelagic species that congregate about them, attracted either by the solid reeflike nature of the supporting structures, or by the numerous smaller forage organisms in the area (bombace et al., 1999, fabi et al., 2002, relini et al., 1976, stanley and wilson, 1991). the reason is that fishes use these artificial structures, for shelter, feeding, spawning, and orientation (kojima, 1956; hunter and mitchell, 1967; gooding and magnuson, 1967; luckhurst and luckhurst, 1978; kakimoto, 1982; ogawa, 1982; steimle and ogren, 1982; yoshimuda, 1982; kellison and sedberry, 1998; rilov and fig. 3. mean number of species and specimens for each combination of levels of factor msc within each level of lsc. bars represent standard errors. no nco mm er cia l u se on ly p. consoli et al.64 benayahu, 1998; caselle et al., 2002; castriota et al., 2011; fabi et al., 2006; leitão et al., 2007). indeed, extractive platforms can furnish shelter for protection from predation, additional food supply and spawning substrate, and can act as a visual attractant for organisms not strictly dependent on hard bottoms (fabi et al., 1998). then, according to these findings, a more complex structure is able to attract more fish species and specimens than a less complex one. in particular, what we observed is that most of the pelagic and demersal fishes were particularly abundant where cross beams and vertical beams cross each other. at these junctions, there is a greater available surface that species such as c. chromis, a. anthias and b. boops use like shelter in case of strong tab. 3. simper of the fish taxa contributing most (%) to the dissimilarity, on large spatial scale, between 4-legs vs 8-legs platforms and, on medium scale, among internal structures with increasing complexity degrees (compl 0, 1 and 2). 4-legs vs 8-legs average dissimilarity=83.26 8-legs 4-legs taxa av. abund. av. abund. contribution % boops boops 5.24 2.84 29.94 anthias anthias 6.1 0 24.29 chromis chromis 3.26 1.29 15.28 spicara flexuosa 1.55 1.7 12.07 thalassoma pavo 1.24 0.53 6.57 oblada melanura 0.52 0.19 4.9 compl 0 compl 1 average dissimilarity=86.06 compl 0 compl 1 taxa av. abund. av. abund. contribution % boops boops 3.2 3.91 28.1 anthias anthias 1.14 4.1 23.82 chromis chromis 0.65 2.25 17.02 spicara flexuosa 1.26 1.36 10.11 thalassoma pavo 0.01 0.93 9.39 oblada melanura 0.98 0.06 5.74 compl 0 compl 2 average dissimilarity=87.07 compl 0 compl 2 taxa av. abund. av. abund. contribution % anthias anthias 1.14 6.96 24.31 boops boops 3.2 6.21 23.41 chromis chromis 0.65 4.9 21.04 thalassoma pavo 0.01 2.07 11.68 spicara flexuosa 1.26 2.19 8.49 oblada melanura 0.98 0.19 4.93 compl 1 compl 2 average dissimilarity=68.01 compl 1 compl 2 taxa av. abund. av. abund. contribution % boops boops 3.91 6.21 25.96 anthias anthias 4.1 6.96 25.08 chromis chromis 2.25 4.9 20.63 thalassoma pavo 0.93 2.07 10.91 spicara flexuosa 1.36 2.19 9.32 av. abund., average abundance. no nco mm er cia l u se on ly the influence of habitat complexity on fish assemblages associated with extractive platforms in the central mediterranean sea 65 currents. c. chromis and a. anthias also use junctions as refuges where to lay eggs: obviously in these places they can better defence the nest from the aggregation of thalassoma pavo specimens, which frequently attacked and destroyed the benthic nests of these two species. moreover, at medium spatial scale, more complex structures provide shelter from predation and current for juvenile and adult fishes: in fact, in case of strong water current many fish species were observed to take refuge on the undercurrent side of these more complex structures. once an industrial decision is made to cease oil and gas production, managers must decide what to do with the structure, a process known as decommissioning and over which a huge debate is animating both scientific communities, stakeholder and common opinion from scientific literature to media (jørgensen et al., 2002; love et al., 2003; schroeder et al., 2004; mangano and sarà, 2017; lucifredi, 2018). the process of decommissioning can be addressed in many ways, from the leaving most part of the structures in place to complete removal. oil and gas platforms have finite economic lives and in the next few decades, several platforms in mediterranean sea will be decommissioned being nearing the end of their economic lives. management decisions regarding the decommissioning of oil and gas platforms will be based on both biological and socioeconomic knowledge baseline (mangano and sarà, 2018), which are essential in evaluating the efficacy of any potential rigs-to-reef program. conclusions the present results could bear strong implications for the environmental management of decommissioned platforms in this basin because the possibility of knowing the attractive potential of an extractive platform could be an important issue in the decommissioning process aiding legislators and resource managers. moreover, further comparative, long-term and at larger spatial scale, studies should be funded in other mediterranean gas/oil platforms, in order to investigate specific cases and propose to maintain a platform rather than another at the end of its life and then lunch a rig-to-reef program enhancing fishery production. apart from the international recommendation on decommissioning options, (i.e. once the topside is removed total removal, partial removal, leave in place; ospar 1982, hamzah 2003) and some case studies from the north seas in a european context (e.g. the indefatigable – inde – field platforms decommissioning project), no specific regulation on decommissioning are prescribed in italy (legislative decree no 257/2016). under the light of the existing literature (mangano and sarà, 2017) future multi-criteria analysis for decommissioning options selection might take into account looking for potential alternative use (e.g. energy production), scientific (e.g. artificial reef monitoring) commercial (e.g. aquaculture, tourism and recreation) and multipurpose (all the above). references aabel jp, cripps s, kjeilen g, 1997. oil and gas production structures as artificial reefs, p. 391-404. in: a.c. jensen (ed.), proceedings of the 1st earrn conference, ancona, italy. european artificial reef research. southampton oceanography centre press, southampton. almany gr, 2004a. differential effects of habitat complexity, predators and competitors on abundance of juvenile and adult coral reef fishes. oecologia 141:105-113. almany gr, 2004b. does increased habitat complexity reduce predation and competition in coral reef fish assemblages? oikos 106:275-284. andaloro f, castriota l, ferraro m, romeo t, sarà g, consoli p, 2011. evaluating fish assemblages associated with gas platforms: evidence from visual census techniques and experimental fishing surveys. cienc. mar. 37:1-9. andaloro f, ferraro m, mostarda e, romeo t, consoli p, 2012. assessing the suitability of a remotely operated vehicle (rov) to study the fish community associated with offshore gas platforms in the ionian sea: a comparative analysis with underwater visual censuses (uvcs). helgoland mar. res. 67:241-250. anderson mj, 2001. a new method for non-parametric multivariate analysis of variance. austral ecology 26:32-46. anderson mj, gorley rn, clarke kr, 2008. permanova+ for primer: guide to software and statistical methods. plymouth, uk: primer-e. anderson mj, ter braak cjf, 2003. permutation tests for multifactorial analysis of variance. j. stat. comput. sim. 73: 85-113. bell jd, r galzin, 1984. influence of live coral cover on coral reef fish communities. mar. ecol. prog. ser. 15:265-274. bell ss, mccoy ed, mushinsky hr, 1991. habitat structure: the physical arrangement of objects in space. chapman and hall, london. bombace g, fabi f, rivas g, 1999. effetti sul popolamento ittico indotti da una piattaforma estrattiva dell’alto adriatico: prospettive di gestione delle risorse costiere. biol. mar. medit. 6:64-72. bohnsack ja, johnson dl, ambrose rf 1991. ecology of artificial reef habitats and fishes, p. 61-107. in: w, seaman jr and l.m. sprague (eds.), artificial habitats for marine and freshwater fisheries. academic press, san diego. bohnsack ja, sutherland dl, 1985. artificial reef research: a review with recommendations for future priorities. bull. mar. sci. 37:, 11-39. buckley rm 1982. marine habitat enhancement and urban recreational fishing in washington. mar. fish. rev. 44:28-37. bull as, kendall jj jr, 1994. an indication of the process: offshore platforms as artificial reefs in the gulf of mexico. bull mar sci 55: 1086-1098. caley mj, st. john j, 1996. refuge availability structures assemblages of tropical reef fishes. j. anim. ecol. 65:414-428. caselle je, love ms, fusaro c, schroeder d, 2002. trash or no nco mm er cia l u se on ly p. consoli et al.66 habitat? fish assemblages on offshore oilfield seafloor debris in the santa barbara channel, california. ices j. mar. sci. 59:258-265. castriota c, falautano m, finoia mg, consoli p, pedà c, esposito v, battaglia p, andaloro f. 2011. trophic relationships among scorpaeniform fishes associated with gas platforms. helgoland mar. res. 66:401-411. chandler cr, sanders rm jr, landry am jr, 1985. effects of three substrate variables on two artificial reef fish communities. bull. mar. sci. 37:129-142. chang k, lee sc, shao kt, 1977. evaluation of artificial reef efficiency based on the studies of model reef fish community installed in northern taiwan. bull. inst. zool. acad. sin. 16:23-36. charbonnel e, serre c, ruitton s, harmelin jg, jensen a, 2002. effects of increase habitat complexity on fish assemblages associated with large artificial reef units (french mediterranean coast). ices j. mar. sci. 59:208-213. clarke kr, 1993. non-parametric multivariate analyses of changes in community structure. aust. j. ecol. 18:117-143. capodici f, ciraolo g, cosoli s, maltese a, mangano mc, sarà g, 2018. downscaling hydrodynamics features to depict causes of major productivity of sicilian-maltese area and implications for resource management. sci. total environ. 628:815-825. consoli p, azzurro e, sarà g, ferraro m, andaloro f, 2007. fish diversity associated to gas platforms: evaluation of two underwater visual census. cienc. mar. 33:121-132. consoli p, martino a, romeo t, sinopoli m, perzia p, canese s, vivona p, andaloro f, 2015. the effect of shipwrecks on associated fish assemblages in the central mediterranean sea. j. mar. biol. assoc. uk 95:17-24. consoli p, romeo t, ferraro m, sarà g, andaloro f, 2013. factors affecting fish assemblages associated with gas platforms in the mediterranean sea. j. sea. res. 77:45-52. de luca m, 1999. international report. offshore 6:38-50. driessen pk, 1985. studing “neptune’s gallery”. sea technol. 26:34-38. fabi g, camilletti e, ciccotti e, luccarini f, lucchetti a, panfili m, solustri c, 1998. ruolo trofico della barriera artificiale di cesano-senigallia nei confronti di alcune specie ittiche. biol. mar. medit. 5:1812-1821. fabi g, grati f, lucchetti a, trovarelli l, 2002. evolution of the fish assemblage around a gas platform in the northern adriatic sea. ices j. mar. sci. 59:309-315. fabi g, grati f, puletti m, scarcella g 2004. effects on fish community induced by installation of two gas platforms in the adriatic sea. mar. ecol. prog. ser. 273:187-194. fabi g, manoukian s, spagnolo a, 2006. feeding behavior of three common fishes at an artificial reef in the northern adriatic sea. bull. mar. sci. 78:39-56. fagundes-netto eb, gaelzer lr, coutinho r, zalmon ir, 2011. influence of a shipwreck on a nearshore-reef fish assemblages off the coast of rio de janeiro, brazil. lat. am. j. aquat. res. 39:103-116. gallaway bj, martin lr, howard rl, boland gs, dennis gd, 1981. effects on artificial reef and demersal fish and macrocrustacean communities. mar. sci. 14:237-299. garcia charton ja, pérez ruzafa a, 2008. correlation between habitat structure and a rocky reef fish assemblage in the southwest mediterranean. mar. ecol. 19:111-128. ghisel rg, 1997. fifty years of offshore oil, gas development. hart publications, houston. gooding rm, magnusson jj, 1967. ecological significance of a drifting object to pelagic fishes. pac. sci. 21 486-497. gorham jc, alevizon ws, 1989. habitat complexity and the abundance of juvenile fishes residing on small scale artificial reefs. bull. mar. sci. 44:662-665. gratwicke b, speight mr 2005. the relationship between fish species richness, abundance and habitat complexity in a range of shallow tropical marine habitats. j. fish. biol. 66:650-667. hamzah ba, 2003. international rules on decommissioning of offshore installations: some observations. mar. pol. 27:339-348. harvey bc, white jl, nakamoto rj, 2005. habitat specific biomass, survival, and growth of rainbow trout (oncorhynchus mykiss) during summer in a small coastal stream. can. j. fish. aquat. sci. 62:650-658. hastings rw, ogren lh, mabry mt, 1976. observations on fish fauna associated with offshore platforms in the northeastern gulf of mexico. fish. bull. 74:387-340. higo n, hashi h, takahama i, tabata s, nagashima m, sakono s, kasmimizutara t, yamasaki t 1980. on the fish gathering effect of the artificial reefs ascertained by the diving observation vii at the sea off maskurazak city. mem. faculty fisheries, kagoshima univ. 29:51-63. hixon ma, beets jp, 1989. shelter characteristics and caribbean fish assemblages: experiments with artificial reefs. bull. mar. sci. 44:666-680. hixon ma, beets jp, 1993. predation, prey refuges, and the structure of coral-reef fish assemblages. ecol. monogr. 63:77-101. hixon ma, menge ba, 1991. species diversity: prey refuges modify the interactive effects of predation and competition. theor. popul. biol. 39:178-200. hunter jr, mitchell ct, 1967. association of fishes with flotsam in the offshore waters of central america. fish. bull. 66:13-29. jones gp, 1988. experimental evaluation of the effects of habitat structure and competitive interactions on the juveniles of two coral reef fishes. j. exp. mar. biol. ecol. 123:115-126. jørgensen t, løkkeborg s, soldal av, 2002. residence of fish in the vicinity of a decommissioned oil platform in the north sea. ices j. mar. sci. 59:288-293. kakimoto h, 1982. the stomach contents of species of fish caught in artificial reefs, p. 271-273. in: s.f vik (ed.), japanese artificial reef technology. tech. rep. 604. aquabio, inc., annapolis. kasprzak ra, 1998. use of oil and gas platforms as habitat in louisiana’s artificial reef program. gulf. mex. sci. 16:37-45. kellison gt, sedberry gr, 1998. the effects of artificial reef vertical profiles and hole diameter on fishes off south carolina. bull. mar. sci. 62:763-780. kojima s 1956. fishing for dolphins in the western part of the japan sea. ii. why do fish take shelter under floating materials? bull. jpn. soc. sci. fish. 21:1049-1052. lassau sa, hochuli df, 2004. effects of habitat complexity on ant assemblages. ecography 27:157-164. leitão f, santos mn, monteiro cc, 2007. contribution of artificial reefs to the diet of the white sea bream (diplodus sargus). ices j. mar. sci. 64:473-478. løkkeborg s, humborstad ob, jørgensen t, soldal av, 2002. no nco mm er cia l u se on ly the influence of habitat complexity on fish assemblages associated with extractive platforms in the central mediterranean sea 67 spazio-temporal variations in gillnet catch rates in the vicinity of north sea oil platforms. ices j. mar. sci 59:294-299. love ms, nishimoto mm, 2012. completion of fish assemblage surveys around manmade structures and natural reefs off california. boem ocs study 2012-020, marine science institute, university of california, santa barbara. love ms, nishimoto mm, schroeder dm, 2010. fish assemblages associated with platforms and natural reefs in areas where data are non-existent or limited. boem ocs study 2010-012, marine science institute, university of california, santa barbara. love ms, schroeder dm, nishimoto mm, 2003. the ecological role of oil and gas production platforms and natural outcrops on fishes in southern and central california: a synthesis of information. ocs study mms 2003-032, us department of the interior, us geological survey, biological resources division, seattle. love ms, york a, 2006. the role of bottom crossbeam complexity in influencing the fish assemblages at california oil and gas platforms. fish. bull. 104:542-549. love ms, westphal w, 1990. comparison of fishes taken by a sportfishing party vessel around oil platforms and adjacent natural reefs near santa barbara, california. fish. bull. 88:599-605. lucifredi a, 2018. [addio alle piattaforme].[article in italian]. le scienze 596:76-81. luckhurst be, luckhurst k, 1978. analysis of the influence of the substrate variables on coral reef fish communities. mar. biol. 49:317-323. macarthur rh, levins r, 1967. the limiting similarity, convergence and divergence of coexisting species. am. nat. 101:377-385. macpherson e, 1994. substrate utilization in a mediterranean littoral fish community. mar. ecol. prog. ser. 114:211-218. mangano mc, bottari t, caridi f, porporato emd, rinelli p, spanò n, johnson m, sarà g, 2017. the effectiveness of fish feeding behaviour in mirroring trawling-induced patterns. mar. environ. res. 131:195-204. mangano mc, kaiser mj, porporato emd, lambert gi, spanò n, 2015. trawling disturbance effects on the trophic ecology of two co-generic astropectinid species. mediterr. mar. sci. 16:538-549. mangano mc, sarà g, 2017. collating science-based evidence to inform public opinion on the environmental effects of marine drilling platforms in the mediterranean sea. j. environ. manage. 188:195-202. mcardle bh, anderson mj 2001. fitting multivariate models to community data: a comment on distance-based redundancy analysis. ecology 82:290-297. mcclanahan tr, 1994. kenyan coral reef lagoon: effects of fishing, substrate complexity, and sea urchins. coral reefs 13:231-241. mccormick mi, 1994. comparison of field methods for measuring surface topography and their associations with a tropical reef fish community. mar. ecol. prog. ser. 112: 7-96. minton v, heath sr, 1998. alabama’s artificial reef program: building oases in the desert gulf. mex. sci. 16:105-106. ogawa y, 1982. reef materials and designs: examples of their applications, p. 320-364. in: s.f vik (ed.), japanese artificial reef technology. tech. rep. 604. aquabio, inc., annapolis. öhman mc, rajasuriya a, 1998. relationships between habitat structure and fish communities on coral and sandstone reefs. environ. biol. fishes 53:19-31. ospar commission, 1992. convention for the protection of the marine environment of the north-east atlantic. available from: https://www.ospar.org/convention porporato em, mangano mc, de domenico f, giacobbe s, spanò n, 2014. first observation of pteroeides spinosum (anthozoa: octocorallia) fields in a sicilian coastal zone (central mediterranean sea). mar. biodivers. 44:589-592. pulsipher ag, daniel wb, 2000. onshore disposition of offshore oil and gas platforms: western politics and international standards. ocean coast. manage. 43:973-995. relini g, geraci s, montanari m, romairone v, 1976. [variazioni stagionali del fouling sulle piattaforme off-shore di ravenna e crotone].[article in italian]. boll. pesca 31:227-256. rilov g, benayahu y, 1998. vertical artificial structures as an alternative habitat for coral reef fishes in disturbed environments. mar. environ. res. 45:431-451. rilov g, benayahu y, 2000. fish assemblage on natural vs vertical artificial reefs: the rehabilitation perspective. mar. biol. 36:931-942. rilov g, benayahu y, 2002. rehabilitation of coral reef-fish communities: the importance of artificial-reef relief to recruitment rates. bull. mar. sci. 70:185-197. roberts cm, ormond rfg, 1987. habitat complexity and coral reef fish diversity and abundance on red sea fringing reefs. mar. ecol. prog. ser. 41:1-8. rooker jr, dokken qr, pattengill cv, holt gj, 1997. fish assemblages on artificial and natural reefs in the flower garden banks national marine sanctuary, usa. coral reefs 16:83-92. scarcella g, grati f, fabi g, 2011. temporal and spatial variation of the fish assemblage around a gas platform in the northern adriatic sea, italy. turk. j. fish. aquat. sci. 11:433-444. schroeder dm, love ms 2004. ecological and political issues surrounding decommissioning of offshore oil facilities in the southern california bight. ocean coast. manage. 47:21-48. stanley dr, wilson ca, 1991. factors affecting the abundance of selected fishes near oil and gas platforms in the northern gulf of mexico. fish. bull. 89:149-159. steimle fw, ogren l, 1982. food of fish collected on artificial reefs in the new york bight and off charleston, south carolina. mar. fish. rev. 44:49-52. terlizzi a, bevilacqua s, scuderi d, fiorentino d, guarnieri g, giangrande a, licciano m, felline s, fraschetti s, 2008. effects of offshore platforms on soft-bottom macro-benthic assemblages: a case study in a mediterranean gas field. mar. poll. bull. 56:1303-1309. yoshimuda n, 1982. discussion of installation planning, p. 137165. in: s.f vik (ed.), japanese artificial reef technology. tech. rep. 604. aquabio, inc., annapolis. warfe dm, barmuta la, 2004. habitat structural complexity mediates the foraging success of multiple predator species. oecologia 141:171-178. willis sc, winemiller ko, lopez-fernandez h, 2005. habitat structural complexity and morphological diversity of fish assemblages in a neotropical floodplain river. oecologia 142:284-295. no nco mm er cia l u se on ly layout 1 introduction heavy metals/trace elements are one of the major occurring constituents in sediments. duffus (2002) suggested the densities of heavy metals range from 3.5 to 7.0 gcm–3 (see also gautam et al., 2015). however, since the threshold varies according to the author, it is impossible to come up with a precise consensus value. therefore, any idea of defining “heavy metals” on the basis of density must be abandoned as it yields nothing but confusion (duffus, 2002). an alternative (and theoretically more acceptable) name for this group of elements is “trace elements” but this is less widely used. since heavy metals/trace elements occur naturally in rock-forming and ore minerals, there is a range of normal background concentrations in soil sediments, water and living organisms (alloway and ayres, 1997). pollution causes a high abnormal concentration of metals/elements relative to the normal background levels. therefore, the only presence of the metals/elements is insufficient evidence of pollution. heavy metals/trace elements are considered serious pollutants not only because of their persistence and nondegradability in the environment, but also because most of them have toxic effects on living organisms when they exceed certain concentrations. additionally, aquatic organisms can bio-accumulate, bio-magnificate or bio-transfer metals to concentrations high enough to bring harmful effects (raulinaitis et al. 2012). bottom sediment is the loose sand, clay, silt and other soil particles and organic matter that settle or deposit at the bottom of a water body. sediments are originated from soil erosion (surface erosion in the watershed) or decomposition of plants, animals and microorganisms within the waterbody. wind, water, and ice are erosion agents that help to carry these particles to streams, rivers and lakes (davies and abowei, 2009; ezekiel et al., 2011). sediments are a sink for many pollutants and trace substances of low solubility and low degree of degradability. therefore, pollutants are conserved in sediments over a long period of time according to their chemical persistence, physical, chemical and biochemical characteristics of the sub-strata (adeyemo et al., 2008). a large proportion of heavy metals/trace elements are originated and dispersed into the air and/or directly into rivers, lakes, sea and reservoirs (todorovic et al., 2014) from anthropogenic activities such as industrial processes, mining, automobile emissions, agricultural activities, wastewater discharge, and urban runoff. natural sources contributing to the presence of heavy metals in sediments include weathering and dissolution of minerals, parent rocks, and soils (decena et al., 2018). sediments can therefore act as natural geosorbent and primary sink for pollutants, including heavy advances in oceanography and limnology, 2018; 9(2): 68-78 article doi: 10.4081/aiol.2018.7576 this work is licensed under a creative commons attribution-noncommercial 4.0 international license (cc by-nc 4.0). the heavy metals/trace elements contents of sediments from owalla reservoir, osun state, southwest nigeria adedeji idowu aduwo,* israel funso adeniyi limnology and hydrobiology laboratory, zoology department, obafemi awolowo university, ile-ife, osun state, nigeria *corresponding author: anjiaduwo@gmail.com abstract the heavy metals/trace elements contents of sediment samples from owalla reservoir were analyzed every three months in two annual cycles (march 2011 – february 2013). the main aim was to measure concentrations of selected elements in sediment samples, their variations in space and seasons and the level of pollution and/or contamination. the bottom sediment samples were collected with a van veen grab and the elemental analysis in the laboratory was based on air-dried samples following standard methods. the overall hierarchy of heavy metals/trace elements in the sediments of the reservoir was in the decreasing order of concentrations: fe >mn >as >zn >ni >co >cr >cu >pb >cd. the concentrations of the heavy metals did not follow any definite pattern from the upstream-downstream basin, although most of them (cu, fe, mn, zn, and co) showed significant differences (p<0.05) in their horizontal variations. mn, pb, co, fe, ni, and zn were significantly (p<0.05) higher at the open water region than in the littoral region. all the elements except ni did not show significant seasonal variations (p>0.05). most of the elements in the reservoir sediment have concentrations within the background levels and concentrations defined in environmental regulations and guidelines, except for as and cd. the contamination factors (cf) for most metals (co, cr, cu, fe, mn, ni, pb, and zn) suggested low contamination in the sediments (cf <1.0). conversely, the sediments were moderately contaminated with cd (cf=2.41) and very highly contaminated with as (cf=19.33). key words: anthropogenic; enrichments; pollution; grab; contamination factor; bioaccumulation; bottom sediment.. received: may 2018. accepted: december 2018. no nco mm er cia l u se on ly the heavy metals/trace elements contents of sediments from owalla reservoir, osun state, southwest, nigeria 69 metals/trace elements (podlasinska and szydlowski, 2017; decena et al., 2018). high concentrations of these contaminants in sediments could be remobilized into the aquatic ecosystems through water and sediment interactions resulting in their bioaccumulation in the tissues of various biota, affecting the distribution and density of benthic organisms (decena et al., 2018). polluted aquatic sediments can also cause environmental damage by releasing metals/trace elements into the surrounding waters, directly contaminating aquatic plants and animals. one of the most common methods for assessing pollution levels and quantifying metals and elemental enrichment in bottom sediments is to compare measured concentrations with a background level or with a concentration posing serious environmental risks (e.g. maximum concentration defined in environmental regulations and guidelines) usually referred to as environmental quality standards (raulinaitis et al. 2012). the geochemical background concentration is defined as the concentration of an element reflecting the natural processes not influenced by human activities. the discrimination between the natural and anthropogenic origin of heavy metals/trace elements is a relative measure that can be determined by the composition in the parent material. the geochemical background concentration allows the distinction of polluted areas from unpolluted ones and is useful for assessing the extent of human activities and the fate (i.e. mobilization, migration, and deposition/uptake of substances in the environment) of these elements. furthermore, it also allows recognizing areas with the higher local background because of the occurrence of mineralization (dung et al., 2013). sediment environmental quality standards are the “measuring stick” against which the acceptability of the presence and concentration of substances in sediment is determined. they help us better assess the quality of the sediment and how clean/healthy it could be. the environmental quality standards are based and designed to protect human and environmental health from the toxicological effects of contaminants (bcmeccs, 2017). the aim of this study is to i) determine the concentrations of selected heavy metals/trace elements in sediment samples from owalla reservoir, ii) their variations in space and seasons, iii) the level of anthropogenic impact in the catchment basin, and iv) heavy metal/trace elements contamination, if any, in the reservoir. methods study area owalla reservoir is located within latitudes 07° 53.5’ and 07° 59.0’ n and longitudes 004° 31.5’ and 004° 35.0’ e with an average elevation of 336±8 m above the mean sea level (fig. 1). the reservoir is situated in the north of osogbo, osun state capital, about 35 km north of ile-ife, approximately 200 km north-east of lagos, 360 km south-west of abuja the capital city of nigeria and bordered by odo otin, ifelodun, irepodun and orolu local government areas of osun state at the northern, eastern, western and southern parts respectively. the dam wall is about 677 m long and 27.5 m high. the total length of the reservoir is about 12 km from the dam wall to the upstream (i.e. the point at which the major river, erinle river, enters into the reservoir) with a maximum width of 3.5 km. it has an area of about 14.5 km2 and capacity of about 94×106 m3. the climate of the study area is typically tropical (ojo, 1977). the area is underlain by the basement complex of southwestern nigeria (kogbe, 1976; rahaman, 1976; rahaman and ocan, 1978). vegetation is the lowland tropical rainforest vegetation (keay, 1959; agboola, 1979; tijani and onodera, 2009). the soils belong to the ferruginous tropical red soil (laterites) with minimal soil degradation and erosion (smyth and montgomery, 1962; tijani and onodera, 2009). the major source of heavy metals/trace elements around the study area could be originated from natural sources mostly by mineralization of parent rock materials. anthropogenic sources are mainly from domestic and agricultural activities within the catchment basin. according to oladejo and ofoezie (2006) and ugbomoiko and ofoezie (2007), the municipalities surrounding the reservoir (i.e. ilie, oba, bara, onipakiti, kuti, igbokiti, idiroko, eko-ende, and ore, etc.) have a population of less than 5000 persons each, with an annual growth rate of about 3%. the inhabitants are mostly yoruba, many of whom are peasant farmers, fishermen or petty traders in agricultural goods. there is no evidence of heavy industrialization. riparian vegetation of the area includes cultivated crops such as vegetables, cashew, mango, cocoa, kola nut, oranges, oil palm, cassava, yam, maize, banana and plantain. sampling stations and programme a total of twenty (20) sampling stations were established across the entire reservoir, representing three different sections (upstream, mid-basin and downstream) and two zones (littoral and open water) (fig. 1; supplementary tab. 1). sampling was carried out every three months aboard a plank boat powered by an outboard-engine in two annual cycles (march 2011 – february 2013). sample collections/field determination and laboratory analysis a van-veen grab of 0.04 m2 area (0.2 m × 0.2 m) was used to collect the bottom sediments for the elemental analysis. samples were taken and kept in a labeled polyno nco mm er cia l u se on ly a.i. aduwo and i.f. adeniyi70 thene bag for laboratory analysis of the heavy metal/trace elements contents of the sediments and analysis of the elements in the sediments was based on air-dried samples, dried in clean trays at room temperature. the laboratory analyses of the heavy metals/trace elements were done using the atomic absorption spectrophotometry (aas) (apha, 1995; international atomic energy agency, 2003; supplementary tab. 2). adequate and proper quality control and quality assurance (qa/qc) measures were observed both at the field and during the laboratory analysis to ensure accurate results and reliability of the data obtained following the appropriate standard methods and instructions. the data obtained were analysed computing descriptive statistics, analysis of variance-anova, and cluster analysis using microsoft excel and paleontological statistics (past, ver. 3.17) software. cluster analysis was computed using the correlation index and the paired group method. calculation of the contamination factor of the heavy metals/trace elements the contamination factor (cf) is an indicator of sediment contamination used in evaluating pollution in an aquatic environment by a given toxic substance (decena et al., 2018). the contamination factor and the degree of contamination (cd) of the heavy metals/trace elements analyzed for the study were used to determine the contamination status of sediments. the cf values for each metal/element based on their overall average values from all the stations in the reservoir were calculated according to the equation given below: contamination factor (cf) = metal/element content of sediment background value of the metal/element where the background value of the metal/element =world surface rock average (geochemical background in the earth crust). cf values for describing the contaminations level of the heavy metals/trace elements in the sediment according to saha and hossain (2011) are as given below: contamination factor level of contamination cf <1 low contamination 1 ≤ cf <3 moderate contamination 3 ≤ cf <6 considerable contamination cf >6 very high contamination. fig. 1. a) lake map and (numbers from 1 to 20) sampling stations. upper reach: stations 1-4; mid-basin: 5-11, 18; lower reach: 12-17, 19-20. littoral, 3-4, 7-9, 11, 13-15, 17-20; open water, 1-2, 5-6, 10, 12, 16. b) map of the study area. no nco mm er cia l u se on ly the heavy metals/trace elements contents of sediments from owalla reservoir, osun state, southwest, nigeria 71 the degree of contamination (cd) is defined as the sum of all contamination factors (cf) for all the metals studied in the given waterbody. results overall variations in owalla reservoir the concentrations of the individual heavy metals/trace elements in the reservoir occurred over the wide range of 0.00 – 377.41 µg g–1 (tab. 1). based on the mean values, they could be grouped into two categories, i.e.: <100 µg g–1=as >zn >ni >co>cr >cu >pb >cd >100 µg g–1=fe >mn comparison with other reservoirs and lakes heavy metals/trace elements in sediments recorded in the present study showed marked similarities with the mass concentrations and hierarchy obtained in some reservoirs in nigeria and other parts of the world (tab. 2). contamination factor and comparison with toxicological reference values in tab. 3, contamination factors (cf) were computed as the ratio between the concentrations of trace elements in the owalla reservoir and the corresponding average concentrations in the earth crust. the degree of contamination, cd, was 22.72. the maximum cf, 19.33, was estimated for cadmium and the least, 0.003, for iron. excluding as, the remaining elements, i.e. co, cr, cu, fe, mn, ni, pb, and zn had cf <1 (table 3). most of the heavy metals/trace elements in the sediments showed concentrations within the background levels and guidelines, with the exception of as and cd, which concentrations were greater than the geochemical background level. moreover, as concentration was greater than the canadian guidelines. cd value was greater than the values for all the guidelines (i.e. uk-cefas, dutch standard, canadian guidelines, and us environmental protection agency). spatial variations the co, cr and fe contents of the reservoir sediment had highest mean values upstream and decreased through the middle portion of the reservoir and towards the dam site (table 4). the lowest values were recorded at the downstream basin. mn showed lowest mean values upstream and increased through the middle portion of the reservoir and downstream. as, cd, cu, ni, pb, and zn showed highest mean values at the middle area of the reservoir; cd and ni had lowest mean values upstream, whereas as, cu, pb, and zn had lowest values downstream. co, cu, fe, mn, and zn showed significant differences (p≤0.05) in their mean values along the (upstream-downstream) reaches of the reservoir, whereas the other metals/elements were not significantly different (p>0.05) (tab. 4). all the heavy/trace metals analysed from the sediment samples collected throughout the study period showed higher mean values at the open water region than in the littoral region of the reservoir (tab. 5). differences were significant (p<0.05) for co, fe, mn, ni, pb, and zn (tab. 5). the relationships among the investigated sampling stations based on the concentrations of the metals/elements in the sediments have been analysed using a cluster analysis (fig. 2). heavy metal concentrations grouped the stations into five clusters. nevertheless, the grouping of stations did not follow any definite and congruent spatial pattern, confirming the absence of any spatial patterns also in the concentrations of the heavy metals/trace elements. seasonal variations as, cd, cu, ni, pb, and zn had higher mean values during the rainy season than the dry season. on the contrary, tab. 1. descriptive statistics of the concentrations of heavy metals/trace elements from the owalla reservoir sediment samples. descriptive statistics s/n parameter unit n min max mean median sem sd skewness kurtosis 1 arsenic (as) µg g–1 152 0.07 37.16 12.06 11.40 0.48 5.91 1.31 5.14 2 cadmium (cd) µg g–1 152 1.85 3.44 2.90 2.91 0.02 0.23 -2.17 8.67 3 cobalt (co) µg g–1 152 2.77 20.78 6.81 6.30 0.24 2.90 1.50 3.38 4 chromium (cr) µg g–1 152 1.47 8.63 3.88 3.39 0.15 1.85 0.46 -0.98 5 copper (cu) µg g–1 152 0.03 32.20 3.19 2.33 0.27 3.39 4.37 33.02 6 iron (fe) µg g–1 152 41.99 377.41 176.22 181.51 4.40 54.27 0.26 1.02 7 manganese (mn) µg g–1 152 23.05 189.88 120.46 128.55 2.96 36.51 -0.59 -0.40 8 nickel (ni) µg g–1 152 0.50 18.58 7.26 6.77 0.27 3.27 0.96 1.22 9 lead (pb) µg g–1 152 0.00 10.70 3.05 2.18 0.22 2.73 0.89 -0.16 10 zinc (zn) µg g–1 152 0.00 45.51 8.99 4.69 0.72 8.90 1.11 1.00 no nco mm er cia l u se on ly a.i. aduwo and i.f. adeniyi72 co, cr, fe, and mn had higher mean values during the dry season than the rainy season. nevertheless, only ni showed significant seasonal differences (p<0.05) (tab. 6). discussion the similarities in the heavy metals/trace elements content recorded from the present study with the levels obtained for some reservoirs in nigeria and other parts of the world with respect to mass concentrations and hierarchy is an evidence of their natural occurrence in rockforming and ore minerals soil and sediments globally. however, the variations observed in a few elements from the selected reservoirs could be ascribed to the nature of the catchment basin and the underlying bedrock geology of the area and the type of anthropogenic activities around the basin. for example, land use pattern is recognized as a crucial factor influencing the concentrations of these elements in terrestrial and aquatic environments in various regions of the world (szarek-gwiazda et al., 2018). unlike most organic pollutants, such as organo-halides, heavy metals occur naturally in a range of normal background concentrations in soil sediments, water and living organisms (alloway and ayres, 1997). pollution gives rise to an anomalously high concentration of metals relative to the normal background levels. therefore, the presence of metals may not be a sufficient evidence of pollution. however, several countries have set environmental quality standard (eqs) or background levels of metals in sediment. such interim guidelines are basically designed for soil and sediment remediation with a wide range of pollutants including, but not limited to, metals (asibor, 2008). the concentrations of most of the heavy metals/trace elements from the present study were generally lower than the geochemical background level (earth crust) representing the world average concentrations of the metals in sediments. exceptions were arsenic (as) and cadmium (cd) that were present with higher average values in the reservoir. asibor (2008) suggested that one tab. 2. comparison of the heavy metals/trace elements average concentrations (µg g–1) in sediments from reservoirs and lakes in nigeria and other countries around the world, and the present study. reservoir/lake metals s/n name location as cd co cr cu fe mn ni pb zn reference 1 opa nigeria 12.28 2523.87 1137.03 0.32 55.68 nathaniel, 2002 2 opa nigeria 183.00 67468.22 2525.89 140.00 155.67 ogunkoya, 2013 3 aiba nigeria 5.82 42.03 28.00 0.44 11.91 olutona et al., 2012 4 asejire nigeria 0.03 43.50 39873.00 1.52 0.05 71.9 20.80 asibor, 2008 5 awassa ethiopia 5.49 8.27 8.69 20.20 15.70 93.80 yohannes et al., 2013* 6 owabi ghana 0.03 4.03 0.10 0.40 akot and abankwa, 2014 7 nasser egypt 113.60 158.30 59.90 784.00 86.60 119.10 lasheen, 1987 8 nasser egypt 0.18 30.79 21.78 12418.00 279.60 27.56 10.91 35.38 goher et al., 2014 9 mariut egypt 38.0 25600.00 958.00 7.30 94.00 biney et al., 1994**** 10 nubia sudan 131.9 332.6 54.7 1196.00 79.40 138.80 lasheen, 1987 11 victoria tanzania 4.74 1.34 24.13 22.98 105.85 machiwa, 2003 12 masinga kenya 33.06 17.31 525.25 12.71 68.51 nzeve et al., 2014 13 yilong china 86.73 31.40 35.99 53.19 86.82 bai et al., 2011* 14 dongting china 29.71 4.65 88.29 47.48 60.99 185.25 li et al., 2013 15 texoma usa 9.00 30.00 38.00 17.00 10.00 89.00 an et al., 2003* 16 erie usa 3.60 64.90 34.30 98.70 162.60 opfer et al., 2011** 17 manchar pakistan 4.70 20.00 21.00 20.10 18.90 96.60 arain et al., 2008* 18 veeranam india 88.20 94.12 63.61 30.06 180.08 suresh et al., 2012* 19 laguna philippines 16.90 103.00 13.00 20.00 13.50 hallare et al., 2005* 20 karla greece 27.40 298.40 38.30 182.80 34.30 31.20 skordas et al., 2015 21 kariba zimbabwe 0.06 29.30 16.10 9.40 42.40 kishe and machiwa, 2003** 22 itá brazil 6.33 90.27 179.86 105301.94 2265.12 20.57 232.04 bonai et al., 2009 23 wallace australia 2.61 28.00 28.00 162.13 2.58 327.25 51.25 29.88 446.25 birch et al., 2001 24 trasimeno italy 9.00 71.70 22.30 30004.00 1566.00 49.00 20.60 59.10 baudo and muntau, 1986*** 25 european lakes (average) 2.41 142.00 96.00 33300.00 2336.00 66.00 135.00 1082.00 morgantini and peruzzi, 2014 26 world average 90.00 45.00 47000.00 850.00 20.00 95.00 filho et al., 2015 27 owalla nigeria 12.06 2.90 6.81 3.88 3.19 176.22 120.46 7.26 3.05 8.99 this study *cited by skordas et al., 2015; **cited by jahangir et al., 2014; ***cited by morgantini and peruzzi, 2014; ****cited by asibor, 2008. no nco mm er cia l u se on ly the heavy metals/trace elements contents of sediments from owalla reservoir, osun state, southwest, nigeria 73 tab. 3. concentrations of heavy metals/trace elements in owalla reservoir, geochemical background levels and the toxicological reference values of the metals in river and lake sediments. concentrations expressed as µg g–1. this study geochemical background uk-cefasa dutch std.a canadian gdls.a s/n metal value cf earth crusta shaleb al–1 al–2 tv rv tel pel us-epac tlesd 1 as 12.06 2.41 5.00** 20.00 100.00 29.00 55.00 7.24 41.60 5.90 2 cd 2.90 19.33 0.15 0.30 0.40 5.00 0.80 7.50 0.70 4.20 0.60 0.60 3 co 6.81 0.27 25.00* 4 cr 3.88 0.04 100.00 90.00 40.00 400.00 100.00 380.00 52.30 160.00 26.00 37.30 5 cu 3.19 0.06 55.00 40.00 40.00 400.00 35.00 90.00 18.70 108.00 16.00 35.70 6 fe 176.22 0.003 56300.00 46700.00 7 mn 120.46 0.14 850.00 950.00 460.00 8 ni 7.26 0.10 75.00 68.00 20.00 200.00 35.00 45.00 15.90 42.80 16.00 18.00 9 pb 3.05 0.24 12.50 20.00 50.00 500.00 85.00 530.00 30.00 112.00 31.00 35.00 10 zn 8.99 0.13 70.00 95.00 130.00 800.00 140.00 720.00 124.00 271.00 110.00 123.00 uk-cefas, united kingdom centre for environment fisheries and aquaculture science; dutch std., dutch standards; canadian gdls, canadian guidelines; al-1, action level 1 (contaminants below this level are of no concern/unlikely to influence licensing decision); al-2, action level 2 (contaminants above this level are considered unsuitable for the environment); tv, target value (target below which the risk to the environment are considered to be negligible); rv, reference value (maximum allowable level, above which the risks to the environment are unacceptable); tel, threshold effect level (exposure to this level is likely to affect some sensitive species); pel, probable effect level (exposure to this level is likely to cause an adverse effect to a wider range of species); us-epa, united state environmental protection agency; tles, threshold level effect in sediments; aesps, 2015; aturekian and wedepohl, 1961; btaylor, 1964; cus-epa, 1999; dburton, 2002; *faboya et al., 2012; **martin and whitfield, 1983) a,b,ccited by goher et al., 2014; **cited by smedley and kinniburgh, 2002. tab. 4. variations of heavy metals/trace elements concentrations in sediments along the major axis of owalla reservoir. reach anova upstream mid-basin downstream s/n parameter unit mean ± se mean ± se mean ± se f p heavy/trace metal contents of sediment (n=32) (n=90) (n=30) 1 arsenic (as) µg g–1 11.33±0.74 12.64±0.71 11.12±0.86 1.049 0.353 2 cadmium (cd) µg g–1 2.82±0.07 2.92±0.02 2.91±0.02 2.275 0.106 3 cobalt (co) µg g–1 8.02±0.67 6.82±0.28 5.53±0.31 6.070 0.003* 4 chromium (cr) µg g–1 4.34±0.38 3.79±0.18 3.69±0.35 1.275 0.282 5 copper (cu) µg g–1 3.07±0.41 3.66±0.42 1.90±0.30 3.161 0.045* 6 iron (fe) µg g–1 193.48±11.39 175.52±5.02 159.89±10.54 3.065 0.050* 7 manganese (mn) µg g–1 104.92±6.00 123.23±3.90 128.73±6.20 4.085 0.019* 8 nickel (ni) µg g–1 6.69±0.55 7.55±0.34 7.00±0.67 0.939 0.393 9 lead (pb) µg g–1 3.22±0.43 3.24±0.30 2.27±0.49 1.509 0.225 10 zinc (zn) µg g–1 8.79±1.45 10.33±1.02 5.21±1.04 3.863 0.023* degree of freedoms, between groups: 2; within groups: 149; *p≤0.05. tab. 5. horizontal variations of the heavy metals/trace elements concentrations in the sediments across the major axis of owalla reservoir. region anova open-water littoral s/n parameter unit mean ± se mean ± se f p heavy/trace metal contents of sediment (n=56) (n=96) 1 arsenic (as) µg g–1 12.10±0.72 12.04±0.63 0.004 0.953 2 cadmium (cd) µg g–1 2.91±0.04 2.89±0.02 0.506 0.478 3 cobalt (co) µg g–1 8.29±0.46 5.95±0.21 26.780 7.189×10–7* 4 chromium (cr) µg g–1 4.22±0.25 3.69±0.19 2.932 0.089 5 copper (cu) µg g–1 3.52±0.31 3.00±0.40 0.831 0.364 6 iron (fe) µg g–1 197.54±6.28 163.78±5.57 14.950 1.640×10–4* 7 manganese (mn) µg g–1 131.06±4.63 114.27±3.71 7.818 0.006* 8 nickel (ni) µg g–1 8.49±0.52 6.54±0.26 13.660 3.06×10–4* 9 lead (pb) µg g–1 3.81±0.38 2.60±0.26 7.338 0.008* 10 zinc (zn) µg g–1 12.60±1.34 6.89±0.76 15.96 1.01×10–4* degree of freedoms, between groups: 1; within groups: 150; *p≤0.05. no nco mm er cia l u se on ly a.i. aduwo and i.f. adeniyi74 of the main reasons of the low metals concentrations he observed in the sediments of asejire reservoir was the low level of industrialization within the catchment basin. on the other hand, kavitha and kumar (2013) associated the accumulation of heavy metals in the sediments and biota of receiving waterbodies to rapid industrialization, urbanization, modern civilization, economic development and increase in population. according to smedley and kinniburgh (2015), arsenic is a ubiquitous element found in the atmosphere, soils, rocks, natural waters, and organisms. it is mobilized in the environment through a combination of natural processes such as weathering reactions, biological activity, and volcanic emissions as well as through a range of anthropogenic activities. most environmental arsenic problems are the result of mobilization under natural conditions, but man has an important impact through mining activity, combustion of fossil fuels, the use of arsenical pesticides, herbicides and crop desiccants and the use of arsenic as an additive to livestock feed, particularly for poultry. from the study area, some of the likely sources of the elevated levels of arsenic may include natural enrichment. the typical richness of arsenic in the earth’s crust is between approximately 2 and 5 mg kg–1. however, an enriched amount may be found in shale and coal deposits of sedimentary and igneous rocks. arsenic adsorption to mineral surfaces (including fe, mn, and al-rich soil and sediments) act as an important sink. arsenic exists mainly in the atmosphere in various forms such as (as2o) adsorbed on particulate matter, which circulates and is returned to the earth by tab. 6. seasonal variations of heavy metals/trace elements concentrations in the sediments samples of owalla reservoir. season anova dry season rainy season s/n parameter unit n mean ± se mean ± se f p 1 arsenic (as) µg g–1 76 11.66±0.89 12.46±0.36 0.687 0.409 2 cadmium (cd) µg g–1 76 2.86±0.03 2.93±0.02 3.138 0.079 3 cobalt (co) µg g–1 76 6.97±0.36 6.66±0.30 0.419 0.518 4 chromium (cr) µg g–1 76 3.96±0.24 3.81±0.18 0.276 0.600 5 copper (cu) µg g–1 76 2.93±0.28 3.46±0.47 0.931 0.336 6 iron (fe) µg g–1 76 179.93±6.51 172.50±5.94 0.711 0.400 7 manganese (mn) µg g–1 76 123.20±3.90 117.72±4.47 0.855 0.357 8 nickel (ni) µg g–1 76 6.55±0.36 7.98±0.38 7.646 0.006* 9 lead (pb) µg g–1 76 2.92±0.29 3.17±0.33 0.333 0.565 10 zinc (zn) µg g–1 76 8.55±1.05 9.44±1.00 0.372 0.543 degree of freedoms, between groups: 1; within groups: 150; *p≤0.05. fig. 2. cluster analysis of the sampling stations based on the heavy metals/trace element contents of the sediments. no nco mm er cia l u se on ly the heavy metals/trace elements contents of sediments from owalla reservoir, osun state, southwest, nigeria 75 wet or dry deposition and simultaneous oxidation and reconversion of arsenic to non-volatile forms (chatterjee et al., 2017). as shown in this study, fe and mn in the sediments had concentrations about ten times higher than the other elements and both have been found to be naturally associated with arsenic occurrence in sediments. natural sources of cadmium from the area could be originated from parent rock materials, soil, plants, and animal matter sunk into the sediment through one or more natural processes like weathering, erosion, and decomposition in the area. according to mislin and ravera (1986), cadmium is found in varying amounts as a natural component of the surface environment in rocks, overburden and soils, water, air, plant and animal tissues with an average concentration in the earth’s crust reported being between 0.15 µg g–1 and 0.11 µg g–1. anthropogenic sources of arsenic and cadmium from the area could come mainly from domestic and agricultural activities within the catchment basin, considering that subsistence agriculture has been the mainstay of more than 90% of the riparian population. since the majority of the farmers in the area practice subsistence agriculture with no access to mechanized agriculture, the use of agro-chemicals to boost crop productions and protections is usually a common practice. as and cd have been found in various concentrations in agro-chemicals used to boost (fertilizers, manures, etc.) and protect crop production (herbicides, pesticides, fungicides, rodenticides, etc.). for example, since the industrial revolution, the use of arsenic as an insecticide, fungicide, and herbicide gradually peaked in 1950s when it was one of the most common pesticides in use, because it was an inexpensive by-product of the smelting of copper, iron, silver, cobalt, nickel, lead, gold, zinc, manganese, and tin. also, agricultural workers preferred to use arsenic and not lead because of its accumulative poison characteristics over insects and pests (gupta, et al., 2017). likewise, phosphate fertilizers and manures used to boost crop productions can contain varying amounts of cadmium (de boo, 1990; roberts, 2014), representing examples of anthropogenic sources of cadmium in the ecosystems. arsenic is listed by the us environmental protection agency (usepa) as one of the priority pollutants and is listed among the most hazardous substances having a significant potential threat to human health (gupta, et al., 2017). arsenic, through ages, is being persistently regarded as a highprofile poison and was related to several conspicuous murder cases, for instance, in the infamous death of napoleon bonaparte in 1851, which was claimed to be a political murder by some conspiracy theorists (gupta, et al., 2017). the u.s. department of health and the european commission’s institute of health and consumer protection have recently summarized the adverse health effects of cadmium on the kidney, renal cortex, pulmonary, cardiovascular, and musculoskeletal systems; moreover, cd was reported as a human carcinogen. for example, cadmium poisoning known as itai-itai disease, a disease causing softening of the bones and kidney failure, was originally discovered in association with rice cultivation in asia (roberts, 2014). concentrations of the heavy metals/trace elements did not follow any definite pattern from the upstream-downstream basin, although most of them showed significant difference in their horizontal variations. further, the highest mean concentrations of the different metals were obtained in various basins of the reservoir. this could probably be explained by the different inputs of pollutants (domestic and agricultural) coming from local populations located within the different areas of the watershed (upstream, midbasin and downstream). this could also confirm that most of the current anthropogenic inputs of these elements into the sediment of the reservoir were mainly of domestic and agricultural sources. most of the elements with significant horizontal variations are generally regarded as nutritional essential elements (i.e. co, cu, fe, mn, and zn). in this regard, heavy metals/trace elements have been grouped into toxic (as, cd, hg, and pb) and nutritional essential (ca, co, cr, cu, fe, mn, mo, ni, se, and zn) elements by alonso et al. (2004). most heavy metals (co, fe, mn, ni, pb, and zn) had significant higher mean values in the open water region than the littoral region of the reservoir. the occurrence of background concentrations of trace elements in soils and sediments is, besides the lithology, also influenced by their clay and organic matter content. therefore, clay and organic matter content are often used to calculate ‘corrected’ background values for trace metal concentrations in soils and sediments (de-saedeleer et al., 2010). also, according to parizanganeh (2008), the elemental concentration of sediments not only depend on anthropogenic and lithogenic sources, but also upon the textural characteristics, hydrogen ion concentration (ph), organic matter content, mineralogical composition and depositional environments of sediments. it is generally believed that metals are associated with smaller grain-size particles. therefore, in this study, the higher concentrations of the heavy metals in the open water region could be due to the higher clay and organic matter content that characterizes this area compared to the littoral region. moreover, the peripheral shallow areas are subject to fluctuating temperatures and erosion of shore materials through wave action. the result is a coastline region of relatively coarse sediment, especially evident near unprotected shores (cole, 1975). increased acidity tends to dissolve and mobilize heavy metals/trace elements in sediments, causing an increase of concentrations in lake water. the hydrogen ion concentration of sediments in the open water zone is usually higher because this stratum of water is generally characterized by decay rather than the production of organic no nco mm er cia l u se on ly a.i. aduwo and i.f. adeniyi76 matter. the sediments in central zones are fine particles largely made up of materials produced within the lake (cole, 1975). as suggested by parizanganeh (2008), the open water region also represents the depositional environment where most of the suspended materials in the water column are usually deposited, whereas the littoral is the eroding region where the action of water movements drags most suspended particles towards the deep stable and less disturbed region. most of the heavy metals/trace elements from the reservoir did not show significant seasonal difference. this could be explained considering that the concentrations of the heavy metals/trace elements in the sediments from this study were the result of the long-term accumulations of these elements over a long period of time rather than fresh anthropogenic enrichments from influx of allochthonous materials due to high incidences of flooding and erosions during the rainy season. conclusions the overall hierarchy of concentrations of heavy metals/trace elements in the reservoir sediment samples based on the mean values was in the decreasing order: fe >mn >as >zn >ni >co >cr >cu >pb >cd. the concentrations of the heavy metals did not follow any definite common pattern from the upstream-downstream basin, although most of them (cu, fe, mn, zn and co) showed significant differences in their horizontal variations, but with various different spatial gradients and highest values in different areas of the reservoir. these unpredictable patterns are an indication that the populations living along the shores in different areas of the watershed might have produced different amounts and types of pollutants (domestic and agricultural sources) discharged to the reservoir. many heavy metal/trace elements analyzed from the sediment samples (mn, pb, co, fe, ni, and zn) showed higher mean values at the open water region than in the littoral region. except ni, all the heavy metal/trace elements did not show significant seasonal difference in their mean values across the dry and rainy seasons. with the exception of as and cd, most of the elements in the sediments showed concentrations within the background levels defined in environmental regulations and guidelines. cf values of the different metals from the sediment samples indicated a low sediment contamination with co, cr, cu, fe, mn, ni, pb, and zn (cf <1.0). conversely, sediments were moderately contaminated with cd (cf=2.41) and very highly contaminated with as (cf=19.33). acknowledgments the authors wish to acknowledge the contribution of mr. adebayo amubieya and godwin zach for their assistance during the field collection. we are grateful to two anonymous reviewers and the editor for valuable comments and suggestions on an earlier version of the manuscript. references adeyemo ok, adedokun oa, yusuf rk, adeleye ea, 2008. seasonal changes in physicochemical parameters and nutrient load of river sediments in ibadan city, nigeria. glob. nest j. 10:326-336. agboola sa, 1979. an agricultural atlas of nigeria. oxford university press, oxford: 248 pp. akot o, abankwa e, 2014. heavy metals contamination and speciation in sediments of the owabi reservoir. environ. res. j. 8:10-16. alloway bj, ayres dc, 1997. chemical principles of environmental pollution. 2nd ed. chapman and hall, london: 302 pp. alonso ml, montaña fp, miranda m, castillo c, hernandez j, benedito jl, 2004. interactions between toxic (as, cd, hg, and pb) and nutritional essential (ca, co, cr, cu, fe, mn, mo, ni, se, zn) elements in the tissues of cattle from nw spain. biometals 17:389-397. apha, 1995. standard methods for the examination of water and wastewater. 19th ed. apha, awwa, wef, washington: 1220 pages. asibor ig, 2008. the macroinvertebrate fauna and sediment characteristics of asejire reservoir, southwest, nigeria. ph.d. thesis, obafemi awolowo university, ile-ife. bcmeccs (british columbia ministry of environment and climate change strategy), 2017. environmental quality standards. available from: https://www2.gov.bc.ca/gov/content/environment/air-land-water/site-remediation/ guidanceresources/technical guidance birch g, siaka m, owens c, 2001. the source of anthropogenic heavy metals in fluvial sediments of a rural catchment: cox river, australia. water air soil pollut. 126:13-35. bonai nc, souza-franco g m, fogolari o, mocelin djc., dal magro j, 2009. distribution of metals in the sediment of the itá reservoir, brazil. acta limnol. bras. 21245-250. burton ag, 2002. sediment quality criteria in use around the world. limnology 3:65-75. chatterjee s, moogoui r, gupta dk, 2017. arsenic: source, occurrence, cycle, and detection, p. 14-29 in: d.k. gupta and s. chatterjee (eds.), arsenic contamination in the environment-the issues and solutions. springer, cham. cole ga, 1975. textbook of limnology. the c.v. mosby co., saint louis: 283 pp. davies oa, abowei jfn, 2009. sediment quality of lower reaches of okpoka creek, niger delta, nigeria. eur. j. sci. res. 26:437-442. de boo w, 1990. cadmium in agriculture. toxicol. environ. chem. 27:55-63. decena scp, arguilles ms, robel ll, 2018. assessing heavy metal contamination in surface sediments in an urban river in the philippines. pol. j. environ. stud. 27:1983-1995. de-saedeleer v, cappuyns v, de-cooman w, swennen r, 2010. influence of major elements on the heavy metal composition of river sediments. geologica belgica 13:257-268. no nco mm er cia l u se on ly the heavy metals/trace elements contents of sediments from owalla reservoir, osun state, southwest, nigeria 77 duffus jh, 2002. “heavy metal”– a meaningless term? pure appl. chem. 74:793-807. dung ttt, cappuyns v, swennen r, phung nk, 2013. from geochemical background determination to pollution assessment of heavy metals in sediments and soils. rev. environ. sci. bio/technol. 12:335-353. environmental statement for port of southampton (esps), 2015. sediment quality: southampton approach chanel dredge, p. 107-121. in: report r/3742/8/r.2015. marine and environment research, abpmer. ezekiel en, hart ai, abowei jfn, 2011. the sediment physical and chemical characteristics in sombreiro river, niger delta, nigeria. res. j. environ. earth sci. 3:341-349. faboya ol, sojinu os, sonibare oo, 2012. an assessment of heavy metals contamination in surface sediments of the niger delta, nigeria. can. j. pure appl. sci. 6:2169-2174. filho fj, marins rv, de lacerda ld, aguiar je, peres tf, 2015. background values for evaluation of heavy metal contamination in sediments in the parnaíba river delta estuary, ne/brazil. mar. pollut. bull. 91:424-428. gautam rk, sharma sk, mahiya s, chattopadhyaya mc, 2015. metals in aquatic media: transport, toxicity and technologies for remediation, p. 1-24. in: s.k. sharma (ed.), heavy metals in water, presence, removal and safety. royal society of chemistry, cambridge. goher me, farhat hi, abdo mh, salem sg, 2014. metal pollution assessment in the surface sediment of lake nasser, egypt. egyp. j. aquat. res. 40:213-224. gupta dk, tiwari s, razafindrabe bhn, chatterjee s, 2017. arsenic contamination from historical aspects to the present, p.1-10. in: d.k. gupta and s. chatterjee (eds.), arsenic contamination in the environment-the issues and solutions. springer, cham. international atomic energy agency, 2003. collection and preparation of bottom sediment samples for analysis of radionuclides and trace elements. iaea-tecdoc-1360. international atomic energy agency, vienna: 130 pp. jahangir tm, khuhawar my, leghari sm, mahar mt, mahar kp, 2014. water quality and sediment assessment of manchar lake, sindh, pakistan: after effects of the super flood of 2010. arab. j. geosci. 8:3259-3283. kavitha p, kumar sp, 2013. evaluation and sediment quality assessment of two perennial ponds in kanyakumari district, tamil nadu, south india. int. j. res. environ. sci. technol. 3:135-144. keay rwj, 1959. an outline of nigerian vegetation. 3rd ed. government printer, nigeria: 46 pp. kogbe ca, 1976. geology of nigeria. university of life center for advanced studies. elizabethan publishing co., lagos: 436 pp. lasheen mr, 1987. the distribution of trace metals in aswan high dam reservoir and river nile ecosystems, p. 235-254 in: t.c. hutchinson and k.m. meema (eds.), lead, mercury, cadmium, and arsenic in the environment. j. wiley & sons, chichester. li f, huang j, zeng g, yuan x, li x, liang j, wang x, tang x, bai b, 2013. spatial risk assessment and sources identification of heavy metals in surface sediments from dongting lake, middle china. j. geochem. explor. 132:75-83. machiwa jf, 2003. metal concentrations in sediments and fish of lake victoria, near and away from catchments with gold mining activities. tanz. j. sci. 29:43-54. martin jm, whitefield m, 1983. the significance of the river input of chemical elements to the ocean, p. 265-296. in: c.s. wong, e. boyle, k.w. bruland and e.d. goldberg (eds.), trace metals in seawater. plenum press, new york. mislin h, ravera o, 1986. cadmium in the environment. birkhäuser verlag, basel: 144 pp. morgantini n, peruzzi l, 2014. surface sediment quality in trasimeno lake. proceedings of 15th world lake conference, perugia, italy. nathaniel it, 2002. the macro-invertebrate benthic fauna and bottom sediment studies of opa reservoir. m.phil. thesis, zoology department, obafemi awolowo university. nzeve jk, kitur ec, njuguna sg, 2014. determination of heavy metals in sediments of masinga reservoir, kenya. j. environ. earth sci. 4:125-132. ogunkoya oo, 2013. ‘all rivers run into the sea, yet the sea is not full…’ obafemi awolowo university inaugural lecture series 256: 61 pp. ojo o, 1977. the climates of west africa. heinemann educational books, ibadan: 219 pp. oladejo so, ofoezie ie, 2006. unabated schistosomiasis transmission in erinle river dam, osun state, nigeria: evidence of neglect of environmental effects of development projects. trop. medi. int. health 11:843-850. olutona go, aribisala og, akintunde ea, 2012. a study of chemical speciation of metals in the aquatic bottom sediment of aiba reservoir, iwo, nigeria. afr. j. environ. sci. technol. 6:312-321. parizanganeh a, 2008. grain size effects on trace metals in contaminated sediments along the iranian coast of the caspian sea, p. 329-336. in: m. sengupta and r. dalwani (eds.), proceedings of the 12th world lake conference. podlasińska j, szydłowski k, 2017. assessment of heavy metal pollution in bottom sediments of small water reservoirs with different catchment management. pol. acad. sci. 3:987-997. rahaman ma, 1976. a review of the basement geology of south western nigeria, p. 41-58. in: c.a. kogbe (ed.), geology of nigeria. elizabethan publication co., lagos. rahaman ma, ocan o, 1978. on relationships in the precambrian migmatite-gneisses of nigeria. j. mining geol. 15:23-30. raulinaitis m, ignatavičius g, sinkevičius s, oškinis v, 2012. assessment of heavy metal contamination and spatial distribution in surface and subsurface sediment layers in the northern part of lake babrukas. ekologija 58:33-43. roberts tl, 2014. cadmium and phosphorous fertilizers: the issues and the science. procedia engin. 83:52-59. saha pk, hossain md, 2011. assessment of heavy metal contamination and sediment quality in the buriganga river, bangladesh, p. 384-388. proceedings 2nd int. conf. on environmental science and technology. iacsit press, singapore. skordas k, kelepertzis e, kosmidis d, panagiotaki p, vafidis d, 2015. assessment of nutrients and heavy metals in the surface sediments of the artificial lake water reservoir karla, thessaly, greece. environ earth sc. 73:4483–4493. smedley pl, kinniburgh dg, 2002. a review of the source, behaviour and distribution of arsenic in natural waters. appl. geochem. 17:517-568. smedley pl, kinniburgh dg, 2015. source and behaviour of no nco mm er cia l u se on ly a.i. aduwo and i.f. adeniyi78 arsenic in natural waters. british geological survey report: 61 pp. smyth aj, montgomery fr, 1962. soil and land use in central western nigeria. the government printer, ibadan: 265 pp. szarek-gwiazda e, mazurkiewicz-boroń g, gwiazda r, urban j, 2018. chemical variability of water and sediment over time and along a mountain river subjected to natural and human impact. knowl. manag. aquat. ec. 419:1-14. taylor sr, 1964. abundance of chemical elements in the continental crust: a new table. geochim. cosmochim. acta 28:1273-1285. tijani mn, onodera s, 2009. hydrogeochemical assessment of metals contamination in an urban drainage system: a case study of osogbo township, south-west, nigeria. j. water resour. protect. 1:164-173. todorovic zb, randelovic lm, marjanovic jz, todorovic vm, cakic md, cvetkovic og, 2014. the assessment and distribution of heavy metals in surface sediments from the reservoir “barje” (serbia). adv. technol. 3:85-95. turekian, kk, wedepohl kh, 1961. distribution of the elements in some major units of the earth’ crust. geol. soc. am. bull. 72:175-192. ugbomoiko us, ofoezie ie, 2007. multiple infection diagnoses of intestinal helminthiasis in the assessment of health and environmental effect of development projects in nigeria. j. helminthol. 81:227-231. us environmental protection agency, 1999. screening level ecological risk assessment protocol for hazardous waste combustion facilities. appendix e: toxicity reference values. solid waste and emergency response. report epa530d99-001c. available from: www.epa.gov/osw no nco mm er cia l u se on ly layout 1 introduction under the influence of the current global warming (ipcc, 2013), the european alps have experienced an average temperature increase of about 2°c since the beginning of the 20th century, which is twice as much as the global average (eea, 2009). this trend is expected to continue with an average temperature increase of 1.5°c (in comparison with 1961-1990 period) by the first half of 21st century, and up to 3.5°c during the second half (gobiet et al., 2014). the cryosphere reduction is the most evident effect of climate change in high mountain areas. mountain glaciers are highly sensitive to the rising air temperature and their melting rate, estimated as ~2 km3 y–1 in the european alps (haeberli et al., 2017), has increased since the 1980s, causing the loss of ~50% of their ice volume (beniston et al., 2018). at this rate, 80-100% of the glaciers in the alps are predicted to disappear by the end of the century (zemp et al., 2006; slemmons et al., 2013), leading to alterations in hydrological regimes, water chemistry and biodiversity in glacial-fed rivers (milner et al., 2017). mountain permafrost (i.e. soil or rock that remains at or below 0°c for at least two consecutive years, dobinski, 2011), is also an important component of the alpine cryosphere that is threatened by climate change (beniston et al., 2018), and its extent in the european alps is estimated to exceed that of glaciers (about 6200 km2 if considering an index ≥0.5, boeckli et al., 2012). the most common permafrost evidence in the alps are rock glaciers, i.e. landforms composed by a mixture of ice and debris and, if active, slowly creeping downslope (schoeneich et al., 2011). rock glaciers are considered as an important water reserve for mountain areas (jones et al., 2018). at present, glaciers and snowpack represent the main drivers of the alpine hydrology, as their seasonal melting is tightly associated with surface and groundwater flow (jansson et al., 2003), while permafrost thawing represents a secondary source to alpine freshwater (krainer et al., 2007; 2011). in fact, subsurface ice is considered as less climatically sensitive than glaciers (jones et al., 2018) and its thawing rate is estimated to be 10-100 times slower than for glaciers due to the thermal advances in oceanography and limnology, 2018; 9(2): 79-96 article doi: 10.4081/aiol.2018.7929 this work is licensed under a creative commons attribution-noncommercial 4.0 international license (cc by-nc 4.0). diatom diversity in headwaters influenced by permafrost thawing: first evidence from the central italian alps federica rotta,1* leonardo cerasino,2 anna occhipinti-ambrogi,1 michela rogora,3 roberto seppi,1 monica tolotti2 1department of earth and environmental sciences, university of pavia, via s. epifanio 14, pavia; 2department of sustainable agro-ecosystems and bioresources, research and innovation centre, fondazione edmund mach, via e. mach 1, s. michele all’adige; 3cnr water research institute, largo tonolli 50, verbania-pallanza, italy abstract glacier melting and permafrost thawing are the most evident effects of the current climate change that is strongly affecting high mountain areas, including the european alps. as the thawing rate of subsurface ice is lower than for glacier ice, it is expected that, while glaciers retreat, an increasing number of alpine headwaters will become more influenced by permafrost degradation during the 21st century. despite the expected change in the relative importance of glacier and permafrost in determining alpine hydrology, studies addressing effects of permafrost thawing on chemical and, especially, biological features of adjacent surface waters are still scarce. the present study contributes to characterise the epilithic and epiphytic diatom diversity in a set of permafrost-fed headwaters in three subcatchments differing in bedrock lithology of the italian central alps (trentino alto-adige) in relation to water chemistry and habitat features. in addition, it explores chemical and biological differences between permafrost-fed streams and headwaters with no direct contact to permafrost, namely glacier-fed (kryal) and precipitation-/groundwater-fed (rhithral) streams. permafrost-fed waters showed higher electrical conductivity and enhanced ion concentrations than glacierand precipitation-fed waters, while concentration of trace elements (e.g. sr, ni, zn, as) were more irregularly distributed among waters of different origin, though they showed a tendency to reach higher levels in permafrost-fed waters. diatom species richness and diversity were lower in permafrost-fed headwaters, and were principally related to water ph and trace metal concentrations. epiphytic diatom assemblages were more diverse than epilithic ones, independently from the water origin, while differences in species composition were not sufficient to unequivocally identify a typical diatom composition for the different water types considered in this study. key words: diatoms; biodiversity; alpine headwaters; mountain permafrost; rock-glacier; water chemistry. received: november 2018. accepted: december 2018. f. rotta et al.80 buffering provided by the thick debris coating (harris and pedersen, 1998). nevertheless, permafrost thawing and its relative importance for the alpine hydrology is expected to increase during the 21st century due to warmer and prolonged summers and reduced precipitation (haeberli et al., 2017). despite the expected change in the relative importance of the major components of the alpine cryosphere, the understanding of the possible cascading effects of permafrost degradation on water quality and ecology of alpine freshwater is still largely incomplete. insights into the effects of rock glacier thawing on inorganic chemistry of alpine headwaters were provided by some recent case studies (krainer et al., 2007; thies et al., 2007, 2013, 2018; colombo et al., 2018a; engel et al., 2018). compared to waters fed by glacier melt or atmospheric precipitation and groundwater, headwaters emerging from active rock glaciers are commonly characterised by higher electrical conductivity (ec) and concentration of major ions, in particular mg2+, ca2+ and so42– (colombo et al., 2018b). in addition, elevated trace metal concentrations (e.g. al, fe, ni, sr) are often detected in permafrost-fed headwaters, especially in metamorphic watersheds (krainer et al., 2007; thies et al., 2007, 2013; ilyashuk et al., 2014; engel et al., 2018). in general, studies specifically addressing the impact of deglaciation on the biota of alpine streams are scarce (fell et al., 2017; hotaling et al., 2017; brown et al., 2018), and studies on the effects of permafrost thaw on the headwater biota are still occasional. primary producers are particularly poorly investigated, despite benthic algae (especially diatoms) play a key ecological role in alpine freshwaters (where external organic input is usually very scarce) as energy source for consumers (rott et al., 2006). nevertheless, first studies show that epilithic diatom assemblages of permafrost-fed running waters are mainly influenced by decrease in water ph and increase in ec and trace metals that appear to be able to induce significant changes in both species’ composition and diversity (thies et al., 2013; lösch et al., 2015). scope of the present work is to contribute in filling the knowledge gap on diatom diversity in permafrost-fed headwaters of the italian central alps (trentino altoadige) in relation to water chemistry and habitat features. the study also aims at enlarging the perspective from one to several sub-catchments and at comparing species richness and diversity of epilithic and epiphytic diatom assemblages. considering the predicted importance of permafrost in determining hydrology, chemistry and biology of headwater in a near future characterized by disappearing glaciers, this work aims also at exploring differences between permafrost-fed streams and headwaters with no direct contact with permafrost, namely glacier-fed (kryal) and precipitation/groundwater-fed (rhithral) streams. this aspect may be ecologically relevant as waters emerging from active rock-glaciers are considered as potential refugia for the kryal biodiversity (hotaling et al., 2017), and as possible stepping stones for the colonization of areas exposed by glacier retreat by organisms of lower altitudes. methods study sites a set of 16 headwaters of different origin were studied in three sub-catchments of the italian central alps (trentino alto-adige), i.e. sulden/solda (su), la mare (mr) and amola (am) valleys (fig. 1 and tab. 1). the lithology of the studied areas consists almost exclusively of crystalline bedrocks. su and mr valleys, which are located respectively in the northern and southern regions of the ortles-cevedale group and belong to the austroalpine domain of the central alps, are dominated by pre-permian metamorphic rocks, and show prevalence, respectively, of phyllades and granitic orthogneiss, and of micaschists (montrasio et al., 2012). the bedrock of the sampled area in the am valley, which belongs to the southalpine domain, consists of oligocene intrusive tonalite rocks (callegari and brack, 2002). the three sub-catchments include glacial landforms of different size and with different degree of debris cover that ranges from clean glaciers, i.e. la mare and suldner/ solda glaciers (mr and su valleys, respectively) to almost completely debris-covered glaciers, i.e. amola (am valley) and klein angelus/angelo piccolo glaciers (su valley). permafrost landforms are widespread in all the three areas, and mainly include active rock glaciers, and inactive and relict rock glaciers at lower elevations. vegetation is absent or very sparse in the upper range of each studied area, while alpine meadows and sparse shrubs reach 2500-2600 m asl, according to the different orographic features of each sub-catchment. the mean annual precipitation ranges from 800-950 mm y–1 at the floor of the su valley (adler, 2015) to ~1200 mm y–1 in the am valley (baroni et al., 2004). studies conducted in the past three decades indicate that these areas are affected by moderate atmospheric deposition of nitrogen (n) compounds, originating from the emission sources located in the po plain, one of the most densely inhabited and industrialized regions of europe (rogora et al., 2006; marchetto et al., 2014). sampling and laboratory analysis water and diatom samples were collected in late summer 2016 (between late august and early september), under condition of maximum melt water contribution from glaciers and rock glaciers and no snowmelt, in order diatom diversity in headwaters influenced by permafrost thawing: first evidence from the central italian alps 81 to capture the highest diatom biodiversity (rott et al., 2006; uehlinger et al., 2010). headwaters fed by permafrost and glaciers (rg and gl, respectively), were sampled at each studied sub-catchment, while rhithral waters fed by atmospheric precipitation, groundwater or of mixed origin (ref), and with no direct contact with permafrost and glaciers, could be found only in the mr and am valleys (fig. 1 and tab. 1). fig. 1. location of the three sub-catchments surveyed in the italian central alps (uppermost left panel) and detailed maps (other panels) indicating the location of the sampled headwaters within each sub-catchment. yellow bullets indicate the sampling sites at each surveyed headwater. the background orthophotos were downloaded from the websites of the provincia autonoma di trento (https://webgis.provincia.tn.it) and the provincia autonoma di bolzano (http://geocatalogo.retecivica.bz.it/geokatalog). su (a) and su (b), portions of the sulden/solda valley located downstream of the glaciers klein angelus/angelo piccolo and suldner/solda, respectively; mr, la mare valley; am, amola valley. f. rotta et al.82 in the su valley, streams fed by the klein angelus/angelo piccolo (su-gl1) and suldner/solda (su-gl2) glaciers were surveyed, together with two springs emerging from two active rock-glaciers located close to each of these two glaciers (su-rg1 and su-rg2, respectively, fig. 1). as the first rg is part of the periglacial belt around the klein angelus/angelo piccolo glacier, it is partially flushed by glacial melting water flowing underground (brighenti, pers. comm). seven different headwaters were surveyed in the mr valley (fig. 1). on the left side of the valley, active rock glaciers and talus slopes with probable presence of permafrost (carturan et al., 2016) originate small streams that partially flow underground due to the slope morphology. four springs emerging from the lowest margin of rockglaciers or talus slopes (tab. 1) were sampled at altitudes between 2710 and 2890 m asl, while a glacial stream (mr-gl) emerging from la mare glacier was sampled at 2898 m altitude on the right valley side (fig. 1). rhithral waters with no direct contact with permafrost were sampled in the upper part of the valley at 2890 m asl, and at the sub-catchment closing section at 2266 m asl (tab. 1), upstream the confluence with the glacial stream. discontinuous permafrost and active rock glaciers characterize the orographic right side of the am valley (baroni et al., 2004; seppi et al., 2012), were samples were collected from two springs emerging from an active rock glacier (am-rg1 and am-rg3 around 2300 m asl, fig. 1) and from a talus slope (am-rg4 at 2570 m asl) characterized by a very low permafrost index according to boeckli et al. (2012). the glacial stream fed by meltwater of the amola glacier (am-gl), located on the left of the valley, was sampled at 2427 m altitude, while the rhithral stream collecting all non-glacial waters originating from the right side of the valley (am-ref, fig. 1) was sampled at 2027 m asl before the confluence with the glacial stream. each sampled headwater has been described by location (altitude and geographic coordinates) and qualitative variables, such as water discharge and turbidity and presence of vegetation inside and outside the water channel, which were empirically estimated on site (tab. 1). water temperature, electrical conductivity (ec), and turbidity (the latter only for the su and part of mr valley) were measured in the field on site with portable conductivity and turbidity meters (wtw gmbh, weilheim, germany). water turbidity of the other surveyed waters and water discharge were empirically estimated by observation during the sampling; water turbidity was given a value between 0 and 5 (tab. 1). water samples for chemical analyses were collected in acid washed high density polyethylene (hdpe) bottles and tab. 1. location, dominant geology, habitat characteristics of the 16 headwaters surveyed in the present study, and type of diatom samples collected. nr id catc lat-x, long-y altid geol veg turb temp pom co32– dia m a.s.l. in, out ntu, (est) °c % dw % loi 1 su-gl1 a 625341, 5156147 2890 m, ph, go -, 42 (2) 3.0 10.9 0.10 e 2 su-rg1 a 624406, 5156079 2728 m, ph, go +, + 6 (1) 1.1 46.5 0.9 e, m 3 su-gl2 a 622653, 5148781 2726 m, ph, go -, 306 (3) 0.1 7.8 6.6 e 4 su-rg2 a 622736, 5149503 2579 m, ph, go +, 3 (0) 1.0 27.3 1.9 e, m 5 mr-gl n 632904, 5118237 2898 m, ms -, n.a. (4) 6.0 6.7 2.9 e 6 mr-rg1 n 627492, 5144697 2710 m, ms -, n.a. (0) 2.4 25.4 3.9 e 7 mr-rg2 n 627645, 5144701 2717 m, ms +, n.a. (0) 4.4 26.4 1.6 e, m 8 mr-rg3 n 626902, 5144681 2809 m, ms -, n.a. (0) 0.6 21.6 5.36 e 9 mr-rg4 n 628372, 5144671 2890 m, ms +, n.a. (0) 2.1 27.3 1.35 e, m 10 mr-ref1 n 628325, 5144673 2266 m, ms +, + n.a. (0) 10.6 52.5 12.4 e, m 11 mr-ref2 n 628325, 5144673 2898 m, ms +, n.a. (0) 9.4 78.1 1.7 e, m 12 am-gl s 631622, 5119213 2427 i, to -, n.a. (5) 2.3 12.0 1.6 e 13 am-rg1 s 632024, 5118102 2321 i, to + + n.a (0). 0.8 94.1 7.7 e, m 14 am-rg3 s 632024, 5118102 2578 i, to -, n.a. (0) 1.1 5.6 0.6 e 15 am-rg4 s 631571, 5118106 2385 i, to +, n.a. (0) 4.5 41.5 4.8 e, m 16 am-ref s 632904, 5118323 2027 i, to +, + n.a. (0) 8.1 21.7 13.9 e, m id, headwater code; am, amola valley; mr, la mare valley; su, sulden/solda valley; gl, glacier-fed; rg, permafrost-fed; ref, rhithral streams fed by precipitation and/or groundwater; catc, major river catchment to which each sampled headwater belongs; s, sarca river; n, noce river; a, adige river; lat-x and long-y, utm latitude and longitude; respectively; altid, altitude, geol, geology of the watershed; i, intrusive; m, metamorphic; to, tonalite; ms, micaschists; ph, phyllades; go, granitic orthogneiss; veg, vegetation present in (in) or ouside (out) each headwater; turb, water turbidity measured as nephelometric turbidity units (ntu) and estimated (0 to 5) by direct observation (est); n.a., not available; temp, water temperature; pom, particulate organic matter of the epilithic biofilm; dw, dry weight; loi, lost on ignition; dia, diatom samples collected from each studied headwater; e, epilithic; m, epiphytic from submerged moss. diatom diversity in headwaters influenced by permafrost thawing: first evidence from the central italian alps 83 preserved at dark and cool (4°c) until analysis. electrical conductivity (ec) at 20°c, ph, alkalinity, major ions (ca2+, mg2+, na+, k+, so42–, cl–), and nutrients (no3-n, nh4-n, total nitrogen (tn), soluble (p-po4) and total phosphorous (tp), sio2) were determined according to standard methods (apha-awwa-wpcf, 2017) at the fem hydrochemistry lab. water samples for the analyses of trace metals were filtered in the field through 0.2 μm mesh polycarbonate filters, collected in acidified 100 ml hdpe bottles and immediately preserved adding 100 μl of hno3 65% (apat, cnr-irsa, 2003). trace metal concentrations were determined by inductively coupled plasma optical emission spectrometry (icp-oes) with ultrasonic nebulizer at the water chemistry laboratory of the water research institute of verbania, italy (cnrirsa), formerly institute of ecosystem study (cnr-ise). further details on the analytical methods and the quality assurance/quality controls adopted in the laboratory can be found at http://www.idrolab.ise.cnr.it/en/. samples for the analysis of the epilithic diatoms assemblages were collected by brushing the surface of 10 different stones randomly collected along a ~10 m long section of each surveyed headwater, according to european committee for standardization (2014). an aliquot of each epilithic sample was immediately subsampled for the determination of organic and carbonate content according to hauer and lamberti (2011). around 0.5 g of freeze-dried material was heated in a furnace at 550°c for 12 hours to quantify the percent content of particulate organic matter (pom), weighted and heated again at 950°c for 3 hours to determine the co32– content to be used as an estimate of primary production. epiphytic diatoms were yielded, wherever possible, by collecting and washing submerged mosses located in the water flow. epiphytic diatoms were retrieved from the majority of the permafrost-fed and rhithral waters, while water mosses were absent in all the surveyed glacier-fed streams. all diatom samples were preserved in 70% ethanol till the cleaning in hot 30% h2o2, and hcl 37%, and the permanent mounting in naphrax® resin, refractive index of 1.7 (european committee for standardization, 2014). oxidation of samples containing particularly resistant organic matter was obtained by adding k2cr2o7 grain by grain (tolotti, 2001). according to the european committee for standardization (2004), at least 400 valves were identified and enumerated along transects of each permanent slide under a leica dm2500 microscope equipped with differential interference contrast optics and a 100x oilimmersion objective (1000x magnification). on the contrary, the total area of the slides relative to the glacialfed streams was scanned, due to the very low diatoms density. taxon identification was based on standard literature (krammer and lange-bertalot, 1986-2004; krammer, 2000; lange-bertalot, 2001; lange-bertalot et al., 2011, 2017), integrated by literature on recently revised taxa (potapova and hamilton et al., 2007; hlúbikova et al., 2009). data analysis variability of ph, major ions responsible for ec (i.e. ca2+, mg2+ and so42–), and more abundant nutrients (no3n, tp and sio2) in the surveyed waters of different origin was compared with one-way anova tests or anova tests on ranks, after testing the normal distribution of the data through a kolmogorov-smirnov distance test. pairwise multiple comparison was performed following the holm-sidak method (holm, 1979). diatom α-diversity of epilithic and epiphytic assemblages was evaluated based on the shannon index (shannon and weaver, 1949) calculated on number of identified individuals for each taxon. the evenness (pielou, 1975) and hill’s index (hill, 1973) were computed to obtain information on the heterogeneity of the assemblages. other diatom data analyses were based on relative abundance (ra) of the diatom taxa identified. preferences of the identified taxa for ph, total phosphorous concentrations, and moisture were evaluated according to the classification proposed by van dam et al. (1994) for european freshwaters. non-metric multidimensional scaling (nmds; kruskal and wish, 1978) was performed by r 3.2.2 (r core team, 2015) software in order to graphically evaluate differences between diatom assemblages identified on different substrata in the studied sites. nmds was applied to a bray and curtis dissimilarity matrix (legendre and legendre, 1998) based on nontransformed relative abundances of the 42 diatom taxa reaching ra ≥0.5% at least in two samples. these restrictions provided the analysis to be performed on an average 95% (84.7-99.7%) of the total diatom abundance of the different samples. epilithic and epiphytic diatom assemblages were evaluated simultaneously in order to obtain a direct comparison of the two diatom communities. samples from the four glacier-fed streams were excluded from the analyses because of the very low number of valves identified in these samples (see below). stress values of nmds configurations with increasing numbers of dimensions were compared in order to evaluate the stability of the analysis and to identify the minimum number of dimensions that is sufficient to explain samples variability (legendre and legendre, 1998). a vector fitting analysis was applied to the sample scores on the final nmds configuration, and computed using the r function envfit included in the vegan package ver. 2.3-0 (oksanen et al., 2015) in order to identify the maximum correlation between the sample distribution within the nmds ordination space and habitat features f. rotta et al.84 and water physical and chemical characteristics of the headwaters studied. except for water temperature and ph, all environmental and chemical data were log transformed (y=log (x+1)) in order to reduce the variability within the data set. in addition, the relation between sample scores on the nmds dimensions and the abundance of single diatom taxa and cumulative ra of diatoms with different preferences for ph, tp, and moisture was explored by correlation analysis. a negative kolmogorov-smirnov distance test indicated the spearman rank correlation coefficient (ρ) as suitable for exploring the correlation within the data matrix. results water chemistry all permafrost-fed waters were characterised at the sampling moment by low discharge (estimated as ranging from ~ 1 l sec–1 for the majority of rg springs to ~5 l sec–1 for am-rg4 and su-rg1), low temperature (ranging from 0.6°c for springs to 4.4°c for waters emerging from talus slopes, i.e. mr-rg2 and am-rg4), and very low water turbidity (tab. 1). no significant differences emerged between the three studied subcatchments. on the other hand, the glacier-fed streams showed higher discharge (estimated as between 5 and 20 l sec–1), high mineral turbidity and variable water temperature in relation to the distance of the sampling site from the glacier snout (tab. 1). the ph values of the majority of the surveyed waters ranged from acidic to neutral, while both the headwaters sampled downstream the suldner/solda glacier (i.e. sugl2 and su-rg2) showed basic values (supplementary tab. 2). although the one-way anova test did not show significant differences in the median ph values of the three water types, glacier-fed and rhithral waters showed higher variability compared to permafrost-fed waters, whose 50% of ph values ranged between 6.6 and 6.9 (fig. 2). median alkalinity of the permafrost-fed waters (5.9±20.3 μeq l–1) was slightly, but not significantly, higher and more variable than values in glacier-fed and rhithral waters (supplementary tab. 2). in general, permafrost-fed waters showed high ec values and ion concentrations (median ec=155±213 μs cm–1), especially in mr and su valleys (up to 516 μs cm–1 ec in mr-rg4), while all headwaters surveyed in the am valley showed scarce mineralization, with ec values never exceeding 20 μs cm–1 (supplementary tab. 2). the ions ca2+, mg2+ and so42– were the most important contributors to the enhanced ec values recorded in the permafrost-fed waters, where they showed quite variable concentrations (fig. 2). due to this high variability, the anova test on ranks (for nonnormally distributed data) indicated that the median concentrations of these ions in permafrost-fed waters (i.e. 71.0, 7.1 and 21 mg l–1, respectively) were not significantly higher than in glacier-fed (i.e. 3.7, 0.5 and 6.8 mg l–1, respectively) and rhithral streams (i.e. 11.9, 2.9 and 38.8 mg l–1, respectively, fig. 2, supplementary tab. 2). in addition, the difference may have been overlooked, since the power of the performed test (0.105) was below the desired values of 0.800, which implies a lower potential to detect differences where they actually exist. concentrations of other ions (na+, k+, cl–) were quite low and did not show significant differences in the surveyed water types (supplementary tab. 2). concentrations of the trace metals were highly variable, as shown in fig. 3, but only in a few cases they exceeded the european limits for drinking waters (eu 83/1998, supplementary tab. 2). sr, al, zn and fe showed the highest median values (fig. 3), which reached ~30 μg l–1 for sr and al, and a pronounced variability, while other metals, such as ni, mn and ba, showed median values <3 μg l–1 but particularly high concentrations in some water samples (supplementary tab. 2). concentrations of the majority of trace metals were very low in all the waters sampled in the am sub-catchment, while the waters sampled in the mr sub-catchment showed higher concentrations (supplementary tab. 2). permafrost-fed waters were enriched in some metals, particularly sr, zn and ni, which reached the highest concentration in waters of the mr (e.g. up to 315 μg l–1 sr in mr-rg2 and 60 μg l–1 ni in mr-rg3 and) su sub-catchments (e.g. 410 μg l–1 sr, 31 μg l–1 as, and 15 μg l–1 ba in su-rg2). on the contrary, al, mn and fe showed higher concentrations in glacier-fed than in permafrost-fed waters in each subcatchment, and reached their maximum in the glacial stream of am (i.e. 211 μg l–1 al and 250 μg l–1 fe in amgl), and mr catchments (i.e. 255 μg l–1 al and 80 μg l–1 mn in mr-gl). the latter stream showed also the highest zn concentration (i.e. 560 μg l–1, supplementary tab. 2). all surveyed waters were very poor in soluble phosphorus (median=2±1 μg l–1) and ammonia concentrations (median=5±6 μg l–1, up to 27 μg l–1 in the glacier-fed site mr-gl, supplementary tab. 2). all glacier-fed streams were characterised by outstanding concentrations of total phosphorous (≥70 µg l–1), while concentrations of total nitrogen showed a strong decreasing gradient from the southernmost (4.8 mg l–1 in am-gl) to the northernmost glaciers (only 0.08 mg l–1 in su-gl1, supplementary tab. 2). median concentrations of no3-n and sio2 were significantly lower in glacier-fed waters of all sub-catchments than in permafrost-fed and rhithral waters (one-way anova test, p<0.01), while values in the permafrost-fed waters did not significantly differ from those of rhithral waters (fig. 3). the highest no3-n and sio2 concentrations were detected in rg springs in the mr diatom diversity in headwaters influenced by permafrost thawing: first evidence from the central italian alps 85 sub-catchment (i.e. 1.4 mg l–1 and 7.8 mg l–1 in mr-rg3 and mr-rg4, respectively, supplementary tab. 2). the median content of pom of the epilithic diatom samples collected from the surveyed waters was 25.9±25.5% of the dry mass (tab. 1), with glacier-fed samples showing the lowest median content (9.3±2.5 %). permafrost-fed and rhithral waters showed a pronounced variability, with no evident relation to the presence of vegetation in the water bed or banks. carbonate content was in general low (median=2.4±4.1 of the mineral ashes, tab. 1), with lower values (≤2.93%) in the samples from all the glacier-fed and several permafrost-fed waters, and fig. 2. concentration ranges of selected water chemical variables measured in headwaters of different origin in the sub-catchments surveyed in the italian central alps. all variables log-transformed (y=(logx+1)), except ph. complete data are available in supplementary tab. 1. gl, glacier-fed; rg, permafrost-fed; ref, rhithral waters fed by precipitation and/or groundwater with no direct contact with glaciers or permafrost. f. rotta et al.86 values exceeding 10% of the ash content only in the rhithral streams in the am and mr sub-catchments (tab. 1). diatom diversity and ecological classification the taxonomical analysis of the 16 epilithic and 10 epiphytic samples collected in summer 2016 yielded a total of 120 diatom taxa. epilithic diatom assemblages of all surveyed glacier-fed waters were extremely sparse, as only up to 42 valvae belonging to a few taxa (2-8) could be identified despite scanning the whole area of the slides (supplementary tab. 3.) permafrost-fed and rhithral headwaters were characterized by denser diatom assemblages, which included several widespread taxa. the most common diatoms (i.e. reaching ra ≥0.5% in more than 10 samples) were achnanthidium minutissimum var. minutissimum (kützing) czarnecki, diadesmis perpusilla (grunow) mann, diatoma mesodon (ehrenberg) kützing, psammothidium helveticum (hustedt) bukhtiyarova round and p. daonense (lange-bertalot) lange-bertalot, and p. marginulatum (grunow) bukhtiyarova and round (tab. 2). the most abundant taxon was the rheophilous a. minutissimum var. minutissimum, (maximum ra=78% in the rhithral stream a-ref-m, tab. 2) which accounted to >30% of the total diatom abundance in the majority of the samples. rare taxa, which reached ra ≥0.5% in only one sample, are listed in supplementary tab. 4. the epilithic diatom assemblages of the permafrostfed waters sampled in the mr sub-catchment showed lower number of taxa, and lower α-diversity (shannon index <2) and heterogeneity compared to permafrost-fed streams in the other sub-catchments (shannon index >2, tab. 3). the sample collected from am-rg4 showed the highest number of taxa (33), and the highest diversity and heterogeneity (tab. 3). the epilithic diatom assemblages of the major reference streams of am and mr subcatchments were comparably diverse, while number of taxa and diversity were much lower in both the samples from mr-ref2 (tab. 3), due to a very pronounced dominance of p. helveticum (ra=76%, tab. 2). on average, the epiphytic diatom samples showed higher species richness (n =23) and diatom diversity (s.i.=2.5±0.1, tab. 3) in comparison to the epilithic ones (n=16, s.i.=1.9±0.1, tab. 3), with highest number of taxa and shannon index in the am-rg4 sample and, secondarily, in the springs emerging from rock-glaciers in the su valley (tab. 3). according to the ecological classification of van dam et al. (1994), the majority (i.e. ~25%) of the diatom taxa identified prefers circumneutral waters and this category accounted in average to ~69% of the total diatom abundance in all the samples (fig. 4a). in some samples with neutral ph values (e.g. mr-rg4 and su-rg2) circumneutral taxa, such as a. minutissimum sensu lato, d. perpusilla and d. mesodon, were accompanied by higher proportions of alkaliphilus diatoms, such as planothidium lanceolatum (brébisson ex kützing) langebertalot), amphora pediculus (kützing) grunow (fig. 4a), and a. lineare (w. smith). acidophilus taxa (i.e. p. marginulatum) represented the most abundant diatom category only in mr-rg3-e, while ra of the acidobiontic taxa eunotia exigua (brébisson ex kützing) rebenhorst exceeded 30% only in the epiphytic sample from mr-ref2 (fig. 4a), in agreement with its quite acidic water ph (supplementary tab. 2). diatom species found in all glacier-fed waters were mainly circumneutral or alkaliphilous, but this classification has to be considered as only indicative due to the very low number of diatom valvae identified in these samples. despite diatom taxa classified as indifferent to the water trophic status represent only ~11% of all identified taxa, in average they accounted to around 40% of the total diatom abundances in all the samples studied (fig. 4b). according to the very low po4-p concentrations of all the surveyed headwaters, oligotraphentic taxa accounted to an average ra of around 22%. however, some mesotraphentic (e.g. d. mesodon) and even eutraphentic taxa (e.g. p. lanceolatum and a. pediculus) reached high relative abundance in some of the analysed epiphytic samples, such as am-rg1-m, am-rg3-m, and mrrg4-m (fig. 4b). only around 10% of all identified diatom taxa were fig. 3. log-transformed (y =(logx+1)) concentration ranges of the trace metals detected in headwaters of different origin in the sub-catchments surveyed in the italian central alps. complete data are in supplementary tab. 2. gl, glacier-fed; rg, permafrost-fed; ref, rhithral waters fed by precipitation and/or groundwater with no direct contact with glaciers or permafrost. diatom diversity in headwaters influenced by permafrost thawing: first evidence from the central italian alps 87 classified as occurring exclusively or mainly in water, while the majority (i.e. ~25% of the total) are considered as able to live also in places that are not permanently submerged (fig. 4c). these tolerant taxa accounted to an average ra of ~47%, while aerophilous diatoms such as d. perpusilla and p. marginulatum were found especially in permafrost-fed springs with low water discharge (e.g. mr-rg1, mr-rg2, mr-rg3 and amrg3), where they showed ra between around 20 and 80% of the total diatom abundance. in general, the tab. 2. list of the 42 taxa used for the nmds analysis (i.e. present with relative abundance ≥ 0.5% in at least two samples). taxa code nr mra sample ph tp achnanthidium minutissimum var. minutissimum (kützing) czarnecki admi 18 78.14 am-ref-m cm id diadesmis perpusilla (grunow) d.g. mann dper 17 80.78 am-rg3-e cm o diatoma mesodon (ehrenberg) kützing dmes 16 61.75 am-rg3-m cm m psammothidium helveticum (hustedt) bukhtiyarova & round phel 14 76.01 mr-ref2-e al m psammothidium daonense (lange-bertalot) lange-bertalot pdao 14 11.85 mr-rg2-m cm o psammothidium marginulatum (grunow) bukhtiyarova & round pmrg 13 87.93 mr-rg3-e ap o encyonema silesiacum (bleisch) d.g. mann esle 9 13.50 am-rg3-m cm id encyonema minutum (hilse) d.g. mann enmi 8 4.10 su-rg1-m cm n.a. psammothidium subatomoides (hustedt) bukhtiyarova & round psat 7 17.36 am-rg1-e ap om psammothidium bioretii (h. germain) bukhtiyarova & round pbio 7 4.55 am-rg4-m cm m nitzschia perminuta (grunow) m. peragallo nipm 6 3.64 am-rg4-m al om hannaea arcus (ehrenberg) r.m. patrick harc 6 2.83 mr-ref1-e al om meridion circulare (greville) agardh mcir 5 27.88 su-rg2-m al id rossithidium petersenii (hustedt) round & bukhtiyarova rpet 5 15.48 su-rg1-m cm o achnanthidium minutissimum var. jakii (rab.) lange-bertalot & ruppel adjk 5 7.93 mr-ref1-m cm n.a. amphora pediculus (kützing) grunow aped 4 43.51 su-rg2-e al e eunotia exigua (brébisson ex kützing) rabenhorst eexi 4 35.85 mr-ref2-m ac id psammothidium helveticum var. minor (flower & jones) buczkó phmi 4 10.61 mr-ref2-m al m fragilaria rumpens (kützing) g.w.f. carlson frum 4 3.64 am-ref-m cm om fragilaria gracilis østrup fgra 4 3.37 mr-ref1-m cm om achnanthidium lineare w. smith acli 3 38.74 su-rg2-e cm n.a. reimeria sinuata (w. gregory) kociolek & stoermer rsin 3 14.35 su-rg1-e cm m amphora inariensis krammer aina 3 6.97 su-rg2-m n.a. o gomphonema parvulum (kützing) kützing gpar 3 6.01 su-rg2-m cm e eunotia palatina lange-bertalot & krüger eplt 3 3.54 mr-ref2-m n.a. n.a. denticula tenuis kützing dten 3 2.04 su-rg1-m al m psammothidium sacculum (j.r. kartner) bukhtiyarova psac 3 2.00 am-rg4-m n.a. n.a. nitzschia alpina hustedt nzal 3 1.84 am-ref-e n.a. o undetermined pennate unpe 3 1.30 su-rg2-e n.a. n.a. gomphonema pumilum var. rigidum reichardt & lange-bertalot gpri 3 1.15 am-ref-e n.a. id planothidium lanceolatum (brébisson ex kützing) lange-bertalot ptla 2 58.68 mr-rg4-e al e luticola mutica (kützing) d.g. mann lmut 2 4.56 mr-rg2-m cm e fragilaria vaucheriae (kützing) j.b. peterson fvau 2 4.20 mr-rg4-m al e fragilariforma virescens (ralfs) d.m. williams & round fvir 2 3.30 mr-ref1-e cm om eunotia intermedia (krasske ex hustedt) nörpel & lange-bertalot euin 2 3.07 mr-ref1-e ap o adlafia suchlandtii (hustedt) monnier & ector adls 2 3.03 su-rg2-e n.a. n.a. adlafia minuscula (grunow) lange-bertalot adms 2 2.96 mr-rg2-m al o staurosira venter (ehrenberg) cleve & j.d.möller ssve 2 1.76 am-rg4-e al me eunotia bigibba kützing ebig 2 1.21 am-rg1-m ap o cyclotella comensis grunow ccms 2 0.97 mr-rg2-e cm n.a. stephanodiscus minutulus (kützing) cleve & möller stmi 2 0.75 am-rg3-m ab he craticula molestiformis (hustedt) mayama cmlf 2 0.61 su-rg1-m al e code, taxon code according to lecointe et al. (1993); nr, number of sample where each taxon occurred with relative abundance > 0.5%; mra, maximum relative abundance reached by each taxon in the sample set; sample, diatom sample code for each surveyed headwater (as in tab. 1); e, epilithic; m, epiphytic; ph and tp, preference for each diatom taxon for water ph and total phosphorus level, respectively, according to the classification proposed by van dam et al. (1994); ac, acidobiontic; ap, acidophilous; cm, circumneutral; al, alkaliphilous; ab, alkalibiontic; o, oligotraphentic; om, oligo-mesotraphentic; m, mesotraphentic; me, meso-eutraphentic; e, eutraphentic; he, hypereutraphentic; id, indifferent to p level; n.a., no classification available. f. rotta et al.88 epiphytic samples showed a higher proportion of hydrophilic taxa (e.g. p. daonense, meridion circulare (greville) agardh, encyonema silesiacum (bleisch) mann, d. mesodon) compared to their epilithic counterparts (fig. 4c, tab. 2). diatom diversity and environmental variables due to the very low number of diatom valvae that could be found in the samples collected from the surveyed glacier-fed headwaters, only epilithic and epiphytic samples from permafrost-fed and rhithral headwaters (n=22) were considered for the nmds ordination. a screen-plot analysis of the stress values of nmds conducted with increasing number of dimensions indicated two dimensions (d1 and d2) as sufficient to represent the variability of diatom assemblages in this study. the two-dimensional nmds explained 86.62% of the data variance with a stress value of 0.15, while nmds analyses including more dimensions provided only slight improvements of both parameters. the distribution of the samples within the nmds space shows a partial grouping according to the water origin (i.e. permafrost-fed or rhithral). the samples from the two major rhithral streams (am-ref and mr-ref1) are located close to the axis origin, and are accompanied by several epilithic and epiphytic samples from permafrost-fed headwaters (fig. 5a). samples from the smaller rhithral stream (mr-ref2) and from other permafrost-fed headwaters are instead located far away from the axis origin in the nmds space, in relation to their high relative abundance of almost exclusive diatom taxa. in fact, both samples from the rock glacier-fed spring surg2 showed high ra of a. lineare and a. pediculus, the epilithic sample from am-rg3 and the epiphytic samples from the permafrost-fed waters am-rg1, am-rg3 and mr-rg2 showed particularly high abundance of d. perpusilla and d. mesodon, respectively (fig. 5b). the epilithic sample from the permafrost-fed mr-rg3 showed high proportion of p. marginulatum, while both samples from mr-ref2 were dominated by varieties of p. helveticum and eunotia spp. (fig. 5b and tab. 2). in contrast, all the samples grouped close to the axis origin were characterised by high relative abundances of a. minutissimum sensu lato, and, secondly, by p. daonense, several fragilariaceae, nitzschia spp., and gomphonema pumilum var. rigidum reichardt & lange-bertalot (fig. 5b). the epilithic and epiphytic samples of the two major rhithral waters are located close to each other in the nmds space, while are far from each other for all the permafrostfed waters, with the sole exception represented by tab. 3. species richness and α-diversity of epilithic and epiphytic diatom assemblages found in the headwaters surveyed in the am, mr, and su sub-catchments. sample taxa s.i.±sd even hetr su-rg1-e 16 2.47±0.08 0.62 5.54 su-rg1-m 27 3.03±0.09 0.64 8.18 su-rg2-e 14 2.03±0.08 0.53 4.08 su-rg2-m 23 2.69±0.08 0.60 6.47 mr-rg1-e 9 1.31±0.06 0.41 2.48 mr-rg2-e 10 1.72±0.07 0.52 3.29 mr-rg2-m 17 2.93±0.08 0.72 7.60 mr-rg3-e 8 0.70±0.08 0.23 1.62 mr-rg4-e 14 1.64±0.08 0.43 3.13 mr-rg4-m 16 2.20±0.08 0.55 4.58 mr-ref1-e 33 2.92±0.12 0.58 7.58 mr-ref1-m 33 2.88±0.01 0.57 7.37 mr-ref2-e 8 1.33±0.09 0.44 2.51 mr-ref2-m 11 2.22±0.07 0.64 4.64 am-rg1-e 8 1.23±0.08 0.41 2.35 am-rg1-m 16 2.25±0.09 0.56 4.77 am-rg3-e 9 1.18±0.09 0.37 2.26 am-rg3-m 22 2.19±0.11 0.49 4.55 am-rg4-e 33 2.91±0.11 0.58 7.50 am-rg4-m 41 3.17±0.11 0.59 8.98 am-ref-e 31 2.87±0.11 0.58 7.33 am-ref-m 21 1.57±0.11 0.36 2.96 sample codes as in tab. 2. taxa, total number of taxa identified in each sample; s.i., shannon index; even, evenness calculated on shannon index; hetr, hill’s heterogeneity. diatom diversity in headwaters influenced by permafrost thawing: first evidence from the central italian alps 89 am-rg4, and to a lesser extent by mr-rg4 (fig. 5a). on the other hand, no sample grouping is evident according to their location in the different sub-catchments studied. the vector fitting revealed highly significant correlations between periphytic diatom and the water variables associated to water acidity, i.e. ph and alkalinity (tab. 4), and to concentrations of several trace metals. pom content of epilithic samples showed a nearly significant (p<0.06) relation. on the contrary, no significant relation emerged with the variables related to fig. 4. relative abundances of diatom taxa with different preferences for a) water ph, b) total phosphorus concentrations, c) moisture in the samples collected from the headwaters surveyed in the italian central alps, according to the classification proposed by van dam et al. (1994). sample codes as in tab. 3. f. rotta et al.90 the water mineralization level, namely ec and major ions. the arrows of the acidity-associated variables point towards the positive side of nmds dimension d2 (fig. 5a), in agreement with the high ph and alkalinity of the headwater su-rg2, while arrows associated to trace metals points to the negative region of nmds d1, where samples characterized by lower ph and higher content of particulate organic matter in the epilithon are grouped. the location of the majority of the sample around the nmds axis origin agrees with their intermediate chemical conditions (including nutrient level) independently from water type and sub-catchment (fig. 5a). spearman rank correlation coefficients between sample scores on the nmds dimensions and diatom-related variables (tab. 3) confirmed the role of a few taxa (i.e. d. fig. 5. nmds ordination of a) the diatom samples collected from the surveyed headwaters based on relative abundance of epilithic and epiphytic diatom taxa reaching relative abundance ≥0.5% in at least two samples; b) species distributions in the ordination space. the vector fitting is restricted to environmental variables significantly correlated (p≤0.05) with the sample ordination. vector orientation indicates the direction of the most rapid change for each variable, vectors lengths indicate the correlation between the variable and the sample ordination. blue circles and triangles, epilithic and epiphytic samples, respectively, from rhithral headwaters fed by precipitation and/or groundwater; yellow circles and triangles, epilithic and epiphytic samples, respectively, from permafrost-fed headwaters. sample and species codes as in tabs. 2 and 3. tab. 4. significant (p<0.05) correlation coefficients between sample scores on nmds dimensions d1 and d2 and chemical and habitat variables (based on the r vector fitting function, r2), and diatom related variables (based on spearman rank correlation, ρ). variable cc p< dim ph 0.51 ** n.a. alk 0.47 ** n.a. hco3 0.46 ** n.a. po4-p 0.44 * n.a. cu 0.43 * n.a. as 0.46 ** n.a. co 0.33 * n.a. mn 0.37 * n.a. zn 0.26 0.07 n.a. pom 0.23 0.08 n.a. admi -0.58 ** d1 dper -0.53 ** d1 dmes 0.56 ** d2 pdao 0.82 *** d1 phel 0.61 ** d1 eexi 0.45 * d1 eplt 0.47 * d1 ac 0.62 * d1 ap 0.50 * d1 id -0.42 * d1 -0.45 * d2 ow 0.52 * d2 mw -0.67 *** d1 -0.46 * d2 ws 0.87 *** d1 ms -0.52 * d1 cc, correlation coefficient; dim, nmds dimension d1 or d2; n.a., not available. diatom codes as in tab. 2; diatom preferences for ph, tp and moisture as in tab. 2; ow, occurring almost exclusively in water; mw, occurring mainly in water; ws, mainly occurring in water but frequent also on wet substrata; ms, mainly occurring on wet or temporarily dry substrata; ***p<0.001; **p<0.01; *p<0.05. diatom diversity in headwaters influenced by permafrost thawing: first evidence from the central italian alps 91 perpusilla, d. mesodon, p. helveticum, e. exigua, eunotia palatina lange-bertalot and krüger) in determining the separation of several samples from permafrost-fed waters along both nmds d1 and d2, from the most numerous group of samples, which were characterized by the dominance of common taxa, especially a. minutissimum var. minutissimum and p. daonense (both correlated to d1, tab. 4). the correlation analyses revealed also a significant positive correlation between samples scores on d1 and the cumulative abundance of acidophilous and acidobiontic taxa, while taxa indifferent to the water phosphorus level showed a negative correlation with the same nmds dimension (tab. 3). furthermore, samples score along nmds d1 are correlated positively with the abundance of taxa occurring both in water and on moist soil, and negatively with taxa occurring mainly in water or on soil, respectively (tab. 3). sample scores on d2 are correlated positively to taxa occurring only in water and negatively to taxa occurring mainly in water, i.e. with scarce tolerance to desiccation (tab. 3). discussion research on alpine rock glaciers has recently gained attention as part of the mountain cryosphere that is increasingly threatened by the global climate warming (ipcc, 2013). however, the knowledge on the impact of climate change on permafrost thawing processes, and, consequently, on the chemical and biological quality of headwaters emerging from alpine rock glaciers is still scarce (colombo et al., 2018). despite the limit posed by the small number of headwater surveyed, the present study confirms the results obtained by previous investigations on chemical features and biodiversity of permafrost-fed headwaters waters in the central alps (thies et al., 2007, 2013, 2018; ilyashuk et al., 2014; lösch et al., 2015; thaler et al., 2015), and provides new insight on aspects scarcely considered so far. similarly to what observed by the mentioned studies, the permafrost-fed springs and streams surveyed in each of the sub-catchments studied in this investigation showed higher ec values and elevated concentrations of ca2+, mg2+ and so42– in comparison to rhithral headwaters with no contact with permafrost. the highest ec values detected in the present study (~500 μs cm–1) are comparable with those found in other catchments on crystalline bedrock of the central alps (krainer et al., 2011; thies et al., 2007, 2013; lösch et al., 2015) and agree with the increasing observations that waters emerging from rock glaciers are characterized by higher solute concentrations (colombo et al., 2018). on the opposite, the glacier-fed streams sampled in each subcatchment were characterized by low mineralization, as commonly observed in glacial waters of siliceous catchments worldwide (slemmons et al., 2013). all headwaters surveyed in this study were very poor in soluble phosphorus, which likely represents a limiting factor for diatom growth, while concentrations of particulate tp and tn were particularly high in the glacier-fed streams, in agreement with previous studies of kryal waters (hodson et al., 2004; rott et al., 2006). the small number of studied sub-catchments does not allow understanding whether the higher tn concentrations detected in the southernmost headwaters may be due to n deposition, owing to their closeness to the atmospheric sources in the po plain, or whether they are casual or related to geomorphological features. the organic content of epilithic biofilms appears to be independent from both water type and catchment geology, as it is likely related to local factors affecting the algal development, such as water turbidity and turbulence. the studied permafrost-fed waters represent a quite heterogeneous group in comparison to glacier-fed and rhithral waters. in fact, all waters on the intrusive bedrock of the am valley were characterised by lower ec and ion concentrations than the waters on the metamorphic bedrock of the mr and su catchments. this suggests that the chemical fingerprint of waters of different origin becomes less clear when considering catchments with different geology. also the difference between permafrost-fed and rhithral waters is not as neat as observed in other areas of the central alps (thies et al., 2007, 2013), which focused on geologically homogeneous watersheds, where chemical differences due to water origin are maximized. a strong geological control of type and amount of weathering products in headwaters has been outlined also in an extensive study on glaciated watershed in western usa (fegel et al., 2016). the double control exerted by water origin and watershed lithology over the concentrations of solutes in alpine headwaters is evident also when considering trace metals. in fact, concentrations were very variable from site to site, but in general lower in the waters of am catchment, and higher in permafrost-fed waters. this supports the hypothesis that trace metals of geological origin can become concentrated in rock glacier outlets due to a set of physical and chemical processes operating in permafrost, such as enrichment related to higher water residence time (esposito et al., 2016) and different groundwater paths (maclean et al., 1999; frey and mcclelland, 2009), freezing-thaw cycles (williams et al., 2006; thies et al., 2007, 2013), or acid rock drainage (ard) induced by sulphide oxidation on freshly exposed rocks (todd et al., 2012). in addition, the fact that only al, mn and fe reached higher levels in glacier-fed than in permafrost-fed waters in each catchment agrees with recent observations that the mentioned processes are not exclusive of permafrost, but can occur also in other highf. rotta et al.92 altitude systems in relation to climate warming (zaharescu et al., 2016). the present study did not provide further elements on the possible mechanism controlling trace metal enrichment in alpine headwaters, as no relation was found with possible regulating factors, such as water ph or particulate matter. however, this study suggests that, although the measured concentrations of trace metals exceed the european guidelines for drinking waters only in a few cases, they potentially represent a threat at local level (e.g. for alpine huts and cattle) in the predicted future context of reduced glacier mass and atmospheric precipitation. the overall diversity of epilithic diatoms in the surveyed headwaters was somewhat lower with respect to previous studies (thies et al., 2013; lösch et al. 2015), in relation to the low species richness and pronounced dominance by a single or a few taxa in several permafrostfed water, especially in the mr sub-catchment. in general, the rhithral waters showed much higher α-diversity, although the highest number of species and shannon diversity was found in a water classified as permafrostfed (i.e. am-rg4-e). the reason for the high diatom diversity of this sample is not clear, but the fact that this headwater emerges from a talus slope, which may contain no ice according to the probabilistic model for permafrost distribution in the alps (boeckli et al., 2012), might explain the similarity between the diatom composition of am-rg4-e and am-ref-e (see below). glacier-fed waters surveyed in this study were characterized by extremely scarce diatom assemblages, in agreement with previous investigations in the alps, tatra and himalaya (uehlinger et al., 1998; cantonati et al., 2001; rott et al., 2006; kawecka et al., 2012; alber, personal communication), but partially in contrast with results from other mountain regions (gesierich and rott, 2012). in the glacier-fed headwaters surveyed in the present study diatom growth appears to be principally limited by physical constraints typically characterizing kryal waters, such as channel instability, reduction of light penetration and scorching exerted by the high mineral turbidity, which reset the diatom succession and reduce taxonomic richness (bona et al. 2012; fell et al., 2018). when excluding the outliers represented by glacierfed waters, the differences between epilithic diatom communities in permafrost-fed and rhithral waters appears to be reduced in comparison with chemical differences. the multivariate analysis outlined that diatom assemblages of around one half of the permafrost-fed samples are highly comparable to those of their rhithral counterparts. these samples were characterized by intermediate chemical conditions, including a moderate to low trace metal concentrations, and their diatom assemblages were correspondingly dominated by the widespread reophilous and tolerant a. minutissmum var. minutissmum, which was accompanied by a set of subdominant species that are typical of oligotrophic, poorly mineralized, circumneutral to slightly acidic arcticalpine or high altitude freshwaters on crystalline bedrock, such as d. mesodon, d. perpusilla, p. daonense, and several fragilarioid and cymbelloid taxa (cantonati et al., 2001; rott et al., 2006; rimet et al., 2007; gesierich and rott, 2004, 2012). interestingly, diatom assemblages of the two rhithral streams sampled in the am and mr subcatchments presented very similar diatom assemblages, despite the much higher mineralization level of the second stream. however, the two waters showed identical ph, which is known as one of the key drivers for benthic diatoms in alpine headwaters (cantonati et al., 2001, 2012; tolotti, 2001; marchetto et al., 2009), and were characterized by similar habitat conditions (e.g. channel and surrounding vegetation, water turbidity and velocity). this stresses the need, when comparing the biodiversity of alpine headwaters, of a holistic evaluation, that considers, beside water chemistry, also the habitat template and the temporal variability of physical constraints, as suggested by uehlinger et al. (2010), sertić perić et al. (2015) and feret et al. (2017). the epilithic assemblages of the remaining permafrost-fed waters differed from those of rithral samples, mainly in relation to high abundances of single or a few taxa, that in turn appeared to be related to peculiar water characteristics. the samples with the most peculiar diatoms assemblage were separated from the group of “intermediate samples” according first to their ph and alkalinity, and secondly to the concentration of some trace metals. alkaliphilous taxa (such as a. lineare) were dominant in su-rg2-e, which showed also the highest concentration of as, ba and sr, while a set of acidophilous taxa (such as p. helveticum, and eunotia spp.) were dominant at the other ph extreme, i.e. in samples that also showed the highest concentrations of co, cu, mn and al. the sample mr-rg3-e represents an outlier, as, despite its slightly acidic ph, it showed high concentration of several metals, especially ni, and was dominated by the acidobionthic p. marginulatum. these results confirm previous evidences of the relation between diatom species composition of alpine headwaters fed by thawing permafrost and heavy metal concentrations (thies et al., 2013; lösch et al., 2015), although trace metal concentrations in the surveyed waters were not so high as in other catchments studied in the central alps (thies et al., 2007, 2013). unlike in these previous studies, no univocal negative effect of trace metals on diatom diversity could be pointed out, since the reduced values of shannon index and heterogeneity of the samples from permafrost-fed waters appeared to be independent from the trace metal levels. in addition, all the “separated” samples were collected in springs or diatom diversity in headwaters influenced by permafrost thawing: first evidence from the central italian alps 93 streams with low water discharge and velocity, and showed, accordingly, high proportion of taxa that can tolerate periods of desiccation. this again stresses the importance of habitat features in regulating diatom biodiversity of alpine headwaters (rott et al., 2006). compared to previous investigations, the present study provides first information also on the epiphytic diatom assemblages of alpine headwaters fed by permafrost, which, in general, were more diverse than epilithic assemblages in all water types. this higher diversity was not due to the presence of typically epiphytic taxa (such as m. circulare), but to the presence of higher proportion of rare taxa, such as several encyonema, gomphonema and nitzschia species, and denticula tenuis kützing. in addition, the multivariate analysis revealed that epilithic and epiphytic diatom assemblages were very similar in rhithral streams, while they were quite different in all the permafrost-fed headwaters. the sole exception is represented by the am-rg4 samples, which showed several characteristic similar to a rhithral water, as mentioned above, and to a lesser extent by mr-rg4. the difference between epilithic and epiphytic diatoms is not easy to be interpreted based on the chemical and habitat data collected in the present study. however, the aquatic mosses are likely able to provide a favourable physical habitat for diatom growth, and to smooth the effects of chemicals, such as trace metals. these aspects surely deserve to be investigated more in detail in the future. conclusions the present study contributes to the chemical and biological characterisation of high mountain freshwaters in the italian alps, with a particular focus on waters fed by permafrost. the results support previous observations outlining the impact of thawing permafrost on the chemistry of headwaters emerging from active rock glaciers, in particular as regards solute content and trace metal concentrations. however, this study clearly indicates that the lithology of the watershed also plays a crucial role in determining the chemical features of headwaters on different bedrocks. in addition, the present study confirms the pivotal importance of ph in modulating the epilithic and epiphytic diatom assemblages of permafrost-fed waters, as well as the capacity of trace metals to negatively affect diatom diversity and to select tolerant diatom species in these waters. on the contrary, the other chemical variables seem to play a subordinate role, despite the high variability of electrical conductivity and major ion concentrations found in the surveyed waters. the comparison of chemistry and diatom diversity in waters with different origin, i.e. fed by glacial melt water, permafrost thawing, and precipitation/groundwater, provides useful information to predict the possible future changes in the physical chemical and biological features of alpine headwaters in relation to the progressing climate-related glacial retreat. in particular, differently to what predicted for other elements of the biota, especially macroinvertebrates (khamis et al., 2014), the diatom αdiversity of kryal headwaters does not seem to be threatened by the glacier retreat. on the contrary, kryal diatom assemblages are expected to become more diverse thanks to the improving environmental condition (i.e. lower physical and chemical shear stress). on the other side, permafrost-fed waters are characterized by a lower diatom α-diversity of the biota, mainly in relation to their enrichment in trace metals, but possibly also to limiting physical habitat conditions, such as low discharge. this suggests that permafrost-fed waters might play a moderate role as biodiversity refugia (hotaling et al., 2017) for benthic microalgae in a warmer future with reduced precipitations. on the contrary, a homogenization and banalization of the periphytic diatom assemblages at both headwater and regional level (i.e. loss of βand γdiversity) can be predicted for the future, due to both the selection of a few tolerant taxa (fell et al., 2018) and to the reduction of habitat variability. within this context, the wide variability of alpine headwaters in relation to water origin, watershed lithology and individual habitat features, seems to have a larger potential in preserving the high-altitude biodiversity. we advocate an increase in the number of case studies across habitats, seasonal aspects and biological groups, in order to improve the understanding the processes affecting diatom diversity in high altitude waters of different origin, and to provide more robust and generalized assessment of the ecological role of these aquatic habitats in the next decades. ackonwledgements the research was financially supported by the edmund mach foundation. the authors want to thank the technical staff of the hydrochemistry lab at fem, and of the water chemistry lab at the cnr-irsa. a special thank goes to stefano brighenti for his precious help during the field work. references adler s, 2015. [das klima von tirol-südtirol-belluno: 1981– 2010, vergangenheit-gegenwart-zukunft].[report in german]. zentralanstalt für meteorologie und geodynamik, südtirol abteilung brandund zivilschutz, bozen: 102 pp. apat, cnr-irsa, 2003. [determinazione di elementi chimici mediante spettroscopia di emissione con sorgente al plasma (icp-oes)]. in: [metodi analitici per le acque].[manual in german]. cnr-irsa. f. rotta et al.94 apha-awwa-wpcf, 2017. standard method for the examination of water and wastewater, 23rd ed. am. publ. healts ass., washington. baroni c, carton a, seppi r, 2004. distribution and behaviour of rock glaciers in the adamellopresanella massif (italian alps). permafr. periglac. process 15:243-259. beniston m, farinotti d, stoffel m, andreassen lm, coppola e, eckert n, fantini a, giacona f, hauck c, huss m, huwald h, lehning m, lópez-moreno ji, magnusson j, marty c, morán-tejéda e, morin s, naaim m, provenzale a, rabatel a, six d, stötter j, strasser u, terzago s, vincent c, 2018. the european mountain cryosphere: a review of its current state, trends, and future challenges. cryosphere 12:759-794. boeckli l, brenning a, gruber s, noetzli j, 2012. permafrost distribution in the european alps: calculation and evaluation of an index map and summary statistics. cryosphere 6:807-820. bona f, la morgia v, falasco e, 2012. predicting river diatom removal after shear stress induced by ice melting. river res. appl. 28:1289-1298. brown le, khamis k, wilkes m, blaen p, brittain je, carrivick jl, fell s, friberg n, füreder l, gislason gm, hainie s, hannah dm, james whm, lencioni v, olafsson js, robinson ct, saltveit sj, thompson c, milner am, 2018. functional diversity and community assembly of river invertebrates show globally consistent responses to decreasing glacier cover. nat. ecol. evol. 2:325-333. callegari e, brack p, 2002.geological map of the tertiary adamello batholith (northern italy): explanatory notes and legend. mem. sci. geol. 54:19-49. cantonati m, corradini g, juttner i, cox ej, 2001. diatom assemblages in high mountain streams of the alps and the himalaya. nova hedwigia 123:37-61. cantonati m, angeli n, bertuzzi e, spitale d, lange-bertalot h, 2012. diatoms in springs of the alps: spring types, environmental determinants, and substratum. freshw. sci. 31:499-524. carturan l, zuecco g, seppi r, zanoner t, borga m, carton a, dalla fontana g, 2016. catchment-scale permafrost mapping using spring water characteristics. permafr. periglac. process. 27:253-270. colombo n, gruber s, martin m, malandrino m, magnani a, godone d, freppaz m, fratianni s, salerno f, 2018a. rainfall as primary driver of discharge and solute export from rock glaciers: the col d’olen rock glacier in the nw italian alps. sci. total environ. 639:316-330. colombo n, salerno f, gruber s, freppaz m, williams m, fratianni s, giardino m, 2018b. review: impacts of permafrost degradation on inorganic chemistry of surface fresh water. glob. planet. change 162:69-83. council of the european union, 1998. directive on the quality of water intended for human consumption, 98/83/ce. official journal 330:32-54. dobinski w, 2011. permafrost. earth. sci. rev. 108:158-169. engel m, brighenti s, bruno mc, tolotti m, comiti f, 2018. multi-proxy approach for characterizing rock glacier outflows, p.1020-1021. in: p. deline, x. bodin and l. ravanel (eds.), proceedings of the 5th european conference on permafrost, chamonix, france. esposito a, engel m, ciccazzo s, daprà l, penna d, comiti, zerbe s, brusetti l, 2016. spatial and temporal variability of bacterial communities in high alpine water spring sediments. res. microbiol. 167:325-333. european committee for standardization, 2004. european standard 14407. water qualityguidance standard for the identification, enumeration and interpretation of benthic diatom samples from running waters. cen, brussels. european committee for standardization, 2014. european standard 13946.water quality guidance for the routine sampling and preparation of benthic diatoms from rivers and lakes.. cen, brussels. european environment agency, 2009. regional climate change and adaptation: the alps facing the challenge of changing water resources. eea, copenahgen: 143 pp. fegel ts, baron js, fountain ag, johnson gf, hall ek, 2016. the differing biogeochemical and microbial signatures of glaciers and rock glaciers. j. geophys. res. biogeosci. 121:919-932. fell sc, carrivick jl, brown le, 2017. the multitrophic effects of climate change and glacier retreat in mountain rivers. bioscience 67:897-911. fell sc, carrivick jl, kelly mg, füreder l, brown le, 2018. declining glacier cover threatens the biodiversity of alpine river diatom assemblages. glob change biol. 24:5828-5840. feret l, bouchez a, rimet f, 2017. benthic diatom communities in high altitude lakes: a large scale study in the french alps. ann. limnol.-int. j. lim. 53:411-423. frey ke, mcclelland jw, 2009. impacts of permafrost degradation on arctic river biogeochemistry. hydrol. process. 23:169-182. gesierich d, rott e, 2004. benthic algae and mosses from aquatic habitats in the catchment of a glacial stream (rotmoos, ötztal, austria). ber. naturwiss.-med. ver. innsb. 91:7-42. gesierich d, rott e, 2012. is diatom richness responding to catchment glaciation? a case study from canedian headwater streams. j. limnol. 71:e7. gobiet a, kotlarski s, beniston m, heinrich g, rajczak j, stoffe m, 2014. 21st century climate change in the european alps a review. sci. total environ. 493:1138-1151. haeberli w, schaub y, huggel c, 2017. increasing risks related to landslides from degrading permafrost into new lake in deglaciating mountain ranges. geomorphology 293:405-417. harris sa, pedersen de, 1998. thermal regimes beneath coarse blocky materials. permafr. periglac. process. 9:107-120. harris c, arenson lu, christiansen hh, etzelmüller b, frauenfelder r, humlum o, kääb a, gruber s, haeberli w, hölzle m, nötzli j, hauck c, isaksen k, lehning m, phillips m, matsuoka n, murton jb, ross n, seppälä m, springman sm, vonder mühll d, 2009. permafrost and climate in europe: monitoring and modelling thermal, geomorphological and geotechnical responses. earth sci. rev. 92:117-171. hauer fr, lamberti ga, 2011. methods in stream ecology. academic press, london: 896 pp. hill mo, 1973. diverisity and evenness: a unifying notation and its consequences. ecology 54:577-586. hodson a, mumford p, lister d, 2004. suspended sediment and phosphorous in proglacial rivers: bioavailability and diatom diversity in headwaters influenced by permafrost thawing: first evidence from the central italian alps 95 potential impacts upon the p status of ice-marginal receiving waters. hydrol. process. 18:2409-2422. holm s, 1979. a simple sequentially rejective multiple test procedure. scan. j. stat. 6:65-70. hotaling s, finn ds, giersch jj, weisrock dw, jacobsen d, 2017. climate change and alpine stream biology: progress, challenges, and opportunities for the future. biol. rev. 92:2024-2045. hlúbikova d, blanco s, falasco e, gomà j. hoffmann l., ector l, 2009. nitzschia alicae sp. nov. and n. puriformis sp. nov., new diatoms from european rivers and comparison with the type material of n. sublinearis and n. pura. j. phycol. 45:742-760. ilyashuk bp, ilyashuk ea, psenner r, tessadri r, koinig ka, 2014. rock glacier outflows may adversely affect lakes: lessons from the past and present of two neighboring water bodies in a crystalline-rock watershed. environ. sci. technol. 48:6192-6200. ipcc, 2013. climate change 2013. in: t.f. stocker, d. qin, g.k. plattner, m. tignor, s.k. allen, j. boschung, a. nauels, y. xia, v. bex and p.m. midgley (eds.). the physical science basis. contribution of working group i to the fifth assessment report of the intergovernmental panel on climate change. cambridge university press, cambridge: 1552 pp. jansson p, hock r, schneider t, 2003. the concept of glacier storage: a review. j. hydrol. 282:116-129. jones db, harrison s, anderson k, betts ra, 2018. mountain rock glaciers contain globally significant water stores. sci. rep. 8:2834. kawecka b, 2012. diatom diversity in streams of the tatra national park (poland) as indicators of environmental conditions. in: a. witkowski (ed.), diatom monographs. gantner verlag, ruggell: 213 pp. khamis kd, hannah m, brown le, tiberti r, milner am, 2014. the use of invertebrates as indicators of environmental change in alpine rivers and lakes. sci. total environ. 493:1242-1254. krainer k, mostler w, spotl c, 2007. discharge from active rock glaciers, austrian alps: a stable isotopeapproach. aust. j. earth sci. 100:102-112. krainer k, chinellato g, tonidandel d, lang k, 2011. analysis of the contribution of permafrost ice to the hydrological water regime. wp7 water resources. report action 7.3, permanet project. available from: http://www.permanetalpinespace.eu/archive/pdf/wp7_3.pdf krammer k, 2000. the genus pinnularia. in: h. lange-bertalot (ed.), diatoms of the european inland water and comparable habitats., gantner verlag, ruggell, germany, 1: 703 pp. krammer k, lange-bertalot h, 1986-2004. süßwasserflora von mitteleuropa. in: h. ettl, j. gerloff, h. heynig, d. mollenhauer (eds.), gustav fischer verlag, stuttgart, germany, bacillariophyceae, naviculaceae, 2/1: 876 pp.; bacillariaceae, epithemiaceae, surirellaceae, 2/2: 596 pp.; centrales, fragilariaceae, eunotiaceae, 2/3: 576 pp.; achnanthaceae, 2/4: 468 pp. kruskal jb, wish m, 1978. multidimensional scaling. sage publications, beverly hills: 93 pp. lange-bertalot h, 2001. navicula sensu stricto, 10 genera separated from navicula sensu stricto, frustulia. in: h. lange-bertalot (ed.), diatoms of the european inland water and comparable habitats. gantner verlag, ruggell: 526 pp. lange-bertalot h, hofmann g, werum m, cantonati m, 2017. freshwater benthic diatoms of central europe: over 800 common species used in ecological assessment. in: m. cantonati, m.g. kelly and h. lange-bertalot (eds.), koeltz botanical books. oberreifenberg: 942 pp. lange-bertalot h, malgorzata b, witkowski a, tagliaventi n, 2011. eunotia and some related genera. in: h. langebertalot (ed.), diatoms of the european inland water and comparable habitats. gantner verlag, ruggell. lecointe c, coste m, prygiel j, 1993. “omnidia”: software for taxonomy, calculation of diatom indices and inventories management. hydrobiologia 269:509. legendre p, legendre l, 1998. numerical ecology. elsevier, amsterdam: 300 pp. lösch b, tolotti m, alber r, 2015. [quellen und bäche mit blockgletschereinfluss].[article in german]. geo. alp. 12:163-182. maclean r, oswood mw, irons jg, mcdowell wh, 1999. the effect of permafrost on stream biogeochemistry: a case study of two streams in the alaskan (usa) taiga. biogeochemistry 47:239-267. marchetto a, rogora m, boggero a, musazzi s, lami a, lotter af, tolotti m, thies h, psenner r, massaferro j, barbieri a, 2009. response of alpine lakes to major environmental gradients, as detected through planktonic, benthic and sedimentary assemblages. fundam. appl. limnol. special issues adv. limnol. 62:419-440. marchetto a, arisci s, tartari ga, balestrini r, tait d, 2014. [stato ed evoluzione temporale della composizione chimica delle deposizioni atmosferiche nelle aree forestali della rete conecofor].[article in italian]. forest@ 11:72-85. milner am, khamis k, battin tj, brittain je, barrand ne, füreder l, cauvy-fraunié s, gíslason gm, jacobsen d, hannah dm, hodson aj, hood e, lencioni v, ólafsson js, robinson ct, tranter m, brown le, 2017. glacier shrinkage driving global changes in downstream systems. pnas 114:9770-9778. montrasio a, berra f, cariboni m, ceriani m, deichmann n, ferliga c, gregnanin a, guerra s, guglielmin m, jadoul f, longhin m, mair v, mazzoccola d, sciesa e, zappone a, 2012. [note illustrative della carta geologica d’italia alla scala 1:50.000, foglio 024].[in italian]. ispra, servizio geologico d’italia, bormio: 150 pp. oksanen j, blanchet fg, kindt r, legendre p, minchin pr, o’hara r, simpson gl, solymos p, stevens mhh, wagner h, 2015. package ‘vegan’. community ecology package, ver.2.2-1. pielou ec, 1975. ecological diversity. j. eiley & sons, new york: 165 pp. potapova mg, hamilton pb, 2007. morphological and ecological variation within theachnanthidium minutissimum (bacillariophyceae) species complex. j. phycol. 43:561-575. r core team (2015). a language and environment for statistical computing. r foundation for statistical computing, vienna, austria. http://www.r-project.org/ rimet f, gomà j, cambra j, bertuzzi e, cantonati m, cappelletti c, ciutti f, cordonier a, coste m, delmas f, tison j, tudesque l, vidal h, ector l, 2007. benthic f. rotta et al.96 diatoms in western european streams with altitudes above 800 m: characterisation of the main assemblages and correspondence with ecoregions. diatom res. 22:147-188. rogora m, mosello r, arisci s, brizzio mc, barbieri a, balestrini r., waldner p, schmitt m, stähli m, thimonier a, kalina m, puxbaum h, nickus u, ulrich, probst a, 2006. an overview of atmospheric deposition chemistry over the alps: present status and long-term trends. hydrobiologia 562:17-40. rott e, cantonati m, füreder l, pfister p, 2006. benthic algae in high altitude streams of the alps a neglected component of the aquatic biota. hydrobiologia 562:195-216. schoeneich p, dall’amico m, deline p, zischg a, 2011. hazards related to permafrost and to permafrost degradation. stateof-the-art report 6.2, permanet project. seppi r, carton a, zumiani m, dall’amico m, zampedri g, rigon r, 2012. inventory, distributiom amd topographiv features of rock glaciers in the southern region of the eastern italian alps (trentino). geogr. fis. dinam. quat. 35:185197. sertić perić m, jolidon c, uehlinger u, robinson ct, 2015. long-term ecological patterns of alpine streams: an imprint of glacial legacies. limnol. oceanogr. 60:992-1007. shannon ce, weaver w, 1949. the mathematical theory of comunication. univerity of illinois press, urbana: 132 pp. slemmons kh, saros je, simon k, 2013. the influence of glacial meltewater on alpin acquatic ecosystems: a review. environ. sci. process impact. 15:1794-1806. thaler b, tait d, tolotti m, 2015. [permafrost und seine auswirkungen auf die ökologie von hochgebirgsseen]. [article in german]. geo. alp 12:183-234. thies h, nickus u, mair v, tessadri r, tait d, thaler b, psenner r, 2007. unexpected response of high alpine lake water to climate warming. environ. sci. technol. 41:7424-7429. thies h, nickus u, tolotti m, tessadri r, krainer k, 2013. evidence of rock glacier melt impacts on water chemistry and diatoms in high mountain streams. cold reg. sci. technol. 96:77-85. thies h, nickus u, tessadri r, tropper p, krainer k, 2018. peculiar arsenic, copper, nickel, uranium, and yttrium-rich stone coatings in a high mountain stream in the austrian alps. austrian j. earth sc. 110. doi: 10.17738/ ajes.2017.0012. todd as, manning ah, verplanck pl, crouch c, mcknight dm, dunham r, 2012. climate-change-driven deterioration of water quality in a mineralized watershed. environ. sci. technol. 46:9324-9332. tolotti m, 2001. phytoplankton and littoral epilithic diatoms in high mountain lakes of the adamello-brenta regional park (trentino, italy) and their relationship to trophic status and acidification risk. j. limnol. 60:171-188. doi: 10.4081 /jlimnol.2001.1.171. uehlinger u, zah r, bürgi h, 1998. the val roseg project: temporal and spatial patterns of benthic algae in an alpine stream ecosystem influenced by glacier runoff. hydrology, water resources and ecology in headwaters. proceedings of the headwater ‘98 conference, merano, italy. iahs publ. 248:419-424. uehlinger u, robinson ct, hieber m, zah r, 2010. the physico-chemical habitat template for periphyton in alpine glacial streams under a changing climate. hydrobiologia 657:107-121. van dam h, mertens a, sinkeldam j, 1994. a coded checklist and ecological indicator values of freschwater diatoms from the netherlands. neth. j. aquat. ecol. 28:17-133. williams mw, knauf m, caine n, liu f, verplanck pl, 2006. geochemistry and source waters of rock glacier outflow, colorado front range. permafr. periglac. process. 17: 13-33. zaharescu dg, hooda ps, burghelea ci, polyakov v, palancasoler a, 2016. climate change enhances the mobilization of naturally occurring metals in high altitude environments. sci. total environ. 560-561:73-81. zemp mh, haeberli w, hoelzle m, paul f, 2006. alpine glaciers to disappear within decades? geophys. res. lett. 33:1-4. layout 1 introduction it has been observed that ephippia pigmentation is a plastic trait (gerrish and cacéres, 2003; hansson, 2004) which can be related to a trade-off between visual predation pressure (mellors, 1975; zaret and kerfoot, 1975; reinikainen, 2012) and better protection for eggs against different types of stress (zaret, 1972; hessen, 1996; gerrish, 2001). mellors (1975) documented that planktivorous fish put greater predation pressure on individuals carrying darker ephippia, but there is more to learn about ephippia pigmentation in different lake types, especially varying in predation degree. in this study, we used a set of model lakes with known typology (plante, 1996b) to document ephippia pigmentation variation. the establishment of la mauricie national park of canada (lmnpc) in 1970 allowed the protection of 536.5 km² of canadian shield of great ecological and cultural value. despite the cessation of major human disturbance (e.g., logging and fishing), the extinction of populations of brook trout (salvelinus fontinalis mitchill, 1814) in small lakes observed during the first half of the 20th century (see bertolo et al., 2008), continued after the establishment of the park. many factors might have caused these extinctions, including the introduction of non-native fishes and transient hypoxia/anoxia events due to beaver (castor canadensis kuhl, 1820) damming (bertolo et al., 2008). the creation of lmnpc was in fact followed by a rise of the beaver population into its territory (masson et al., 2001) due to the end of trapping activities (plante, 1996a). whereas the actual occurrence of most fish species, and especially the presence or absence of brook trout, is relatively well known for most lakes in the lmnpc, the picture is less clear for the period preceding 1970. the historical data referring to the period before the creation of the park, which were obtained from fishing advances in oceanography and limnology, 2016; 7(2): 197-205 article doi: 10.4081/aiol.2016.6215 this work is licensed under a creative commons attribution-noncommercial 4.0 international license (cc by-nc 4.0). brook trout (salvelinus fontinalis) extinction in small boreal lakes revealed by ephippia pigmentation: a preliminary analysis alexandre bérubé tellier,1* paul e. drevnick,2# andrea bertolo1 1département des sciences de l'environnement, université du québec à trois-rivières, 3351 bd des forges c.p.500, trois-rivières g9a 5h7, québec, canada; 2institut national de la recherche scientifique, centre eau terre environnement, université du québec 490 de la couronne, québec g1k 9a9, canada #present address: university of michigan, biological station, 440 church st., ann arbor, mi 48109, usa *corresponding author: alexandre.berube.tellier@uqtr.ca abstract ephippium pigmentation is a plastic trait which can be related to a trade-off between visual predation pressure and better protection of cladoceran eggs against different types of stress. experimental studies showed that planktivorous fish exert a greater predation pressure on individuals carrying darker ephippia, but little is known about the variation of ephippium pigmentation along gradients of fish predation pressure in natural conditions. for this study, our sampling design included four small boreal lakes with known fish assemblages. two of the lakes have viable brook trout (salvelinus fontinalis) populations, whereas the other two lakes experienced brook trout extinctions during the 20th century. cladoceran ephippia were extracted from sediment cores at layers corresponding to the documented postextinction phase (1990’s) and from an older layer (1950’s) for which the brook trout population status is not known precisely. our first objective was to determine whether brook trout extinction has a direct effect on both ephippium pigmentation and size. our second objective was to give a preliminary assessment of the status of brook trout populations in the 1950’s by comparing the variation in ephippia traits measured from this layer to those measured in the 1990’s, for which the extinction patterns are well known. cost-effective image analysis was used to assess variation in pigmentation levels in ephippia. this approach provided a proxy for the amount of melanin invested in each ephippium analysed. our study clearly shows that ephippium pigmentation may represent a better indicator of the presence of fish predators than ephippium size, a trait that showed a less clear pattern of variation between lakes with and without fish. for the 1990’s period, ephippia from fishless lakes were darker and showed a slight tendency to be larger than ephippia from lakes with brook trout. however, no clear differences in either ephippium size or pigmentation were observed between the 1990’s and 1950’s layers within each lake. this suggests that brook trout extinction already occurred before the 1950’s, or that brook trout population abundance was already extremely low before the 1990’s. our preliminary study shows that ephippium pigmentation can be used as a tool to quickly assess present and past predation levels on zooplankton when only sediment samples are available. key words: cladocerans; ephippia; extinction; pigmentation; predation; salvelinus fontinalis. received: july 2016. accepted: november 2016. no n c om me rci al us e o nly 198 a. bérubé tellier et al. clubs and other historical archives (plante, 1996b), do not always provide accurate information about the status of the fish community or the timing of extinction events. although the documented presence of viable brook trout populations in some lakes after the creation of lmnpc indicates that the species was also present before, identifying the exact moment of brook trout extinction events in other lakes remains a challenge. here we propose to use a paleolimnological approach to help solve this issue. paleolimnological tools can be used to indirectly infer the presence of planktivorous fish (jeppesen et al., 2002; davidson et al., 2011) at the moment of park creation by providing estimates of key traits of the zooplankton community related to predation pressure. in particular, sediments accumulated at the bottom of lakes contain cladoceran ephippia, which can give an approximate portrait of the size structure of the cladoceran assemblage and allow us to infer about the levels of predation in which they were produced (brooks and dodson, 1965; jeppesen et al., 2002). it has been observed that large cladocerans possess a wide range of phenotypic plasticity and genetic variation associated to body size, that are widely used as indicators of predation pressure from planktivorous fish (jeppesen et al., 2002; dzialowski et al., 2003). it has also been observed that more pigmented individuals are exposed to stronger predation pressure by visual predators (zaret and kerfoot, 1975). daphnia carrying darker ephippia are known to suffer greater predation than specimens with less pigmented ephippia (mellors, 1975). as a result, cladocerans tend to reduce the pigmentation of appendices and organs (zaret, 1972) or to reduce the size of darker body parts in the presence of planktivorous fish (reinikainen, 2012). a similar phenomenon is observable also for copepods, where the presence of carotenoids, while allowing a better protection against both uv radiation and parasites, makes them more conspicuous to predators (hansson, 2004; van der veen, 2005). therefore, in different species of both cladocerans and copepods it is possible to observe a trade-off between resistance to stress factors and vulnerability to visual predators, which involves chemical compounds responsible for pigmentation (van der veen, 2005). the main objective of this study was to determine whether brook trout extinction directly affects both ephippium pigmentation and size in two lakes of lmnpc, while secondarily giving a preliminary assessment of the timing of brook trout extinction in these lakes, by using a paleolimnological approach. cladoceran ephippia from sediment cores collected from two lakes in which brook trout populations went extinct were compared to those collected from two lakes with viable brook trout populations. the size and pigmentation of ephippia produced before the creation of lmnpc were compared in both types of lakes to those produced after its creation. based on historical records, all four study lakes possessed viable brook trout populations during the first half of the 20th century (i.e., before the creation of lmnpc), but intensive sampling conducted by park canada in the 1990’s confirmed the presence of viable brook trout populations in only two of the lakes (plante, 1996b). therefore, although we only have approximate and uncertain information about the timing of brook trout extinction, we have a clear picture of the fish assemblage for the period following the creation of lmnpc. this study was based on the analysis of both pigmentation and biometry of cladoceran ephippia deposited in the sediments during the 1950’s and 1990’s (based on 210pb core dating). two main hypotheses were tested concerning the expected patterns of planktivory between lakes that experienced brook trout extinction and lakes with viable brook trout populations. first, the ephippia produced in the 1990’s should be smaller in lakes with brook trout than in those that experienced its extinction. second, lakes with brook trout are expected to show lower levels of pigmentation in the ephippia produced in the 1990’s than those produced in the other lakes. these expected results should reflect the reduction of predatory pressure related to the extinction of brook trout. in contrast, no clear patterns are expected for the 1950’s period for neither ephippium size nor pigmentation, since we hypothesized that brook trout were present in all of the study lakes at that time and went extinct only after the creation of lmnpc. methods study site the four study lakes (alphonse, genévrier, giron and noir, tab. 1) are located in lmnpc, in the upper part of the mauricie region (46°46’n, 73°00’w, québec, canada). these are headwater lakes, located at an average elevation of 287 m asl, of relative small size (average of 12.2 hectares) and relatively shallow (average depth of 6.7 meters). lmnpc archives suggest that all lakes found on this territory were historically inhabited by brook trout (lacasse and magnan, 1994; plante, 1996b). lakes alphonse and giron still possess viable populations of brook trout that are currently exploited for sport fishing, whereas the brook trout populations of lakes genévrier and noir are now extinct. four introduced fish species (tab. 1) are present in lake giron in addition to brook trout, and brook trout of lake alphonse are in sympatry with a cyprinid fish (tab. 1). no other fish species are present in the two lakes where brook trout extinction has occurred, allowing the comparison of situations with complete absence of fish and situations with documented presence of fish (brook trout alone or accompanied by other no n c om me rci al us e o nly response of ephippium pigmentation to fish 199 planktivorous fishes). lake genévrier suffered a presumed extinction of its population in the 1980’s, with the last brook trout catches by lmnpc staff in 1984 (plante, 1996b; masson et al., 2001). in lake noir, there is no documented historical proof of natural presence of brook trout, but the fishing club’s archives indicate the lake was stocked with brook trout (at least) during the 1960’s (plante, 1996b). sediment sampling and radioisotopic dating short sediment cores were collected with an hth gravity corer (pylonex ab, umeå, sweden) at the deepest point of each lake during the ice-covered period in 2013. from each lake, duplicate sediment cores (approximately 30 cm length) were collected from one or two ice holes made within a 3m² area; one core was used for radioisotopic dating and the second for ephippia extraction and analysis. intact sediment cores were transported to lmnpc laboratories (saint-mathieu or saint-jean-despiles, according to proximity with the sampled lake), where the cores were vertically extruded and sectioned into 1-cm thick slices. the subsampled sections were placed into individual plastics bags and stored in the dark at 4°c until further manipulation. we ensured the parallel cores retrieved from each lake represented replicates of the same stratigraphic intervals by comparing the profiles of organic content obtained by loss on ignition (loi). to achieve this goal, 0.25g of dried sediment (100°c overnight) from each sediment layer were combusted for one hour at 550°c. loi manipulations were performed in two different laboratories, as one core was at the institut national de la recherche scientifique, centre eau terre environnement for dating and the other was at université du québec à trois-rivières for ephippial analysis. respectively, the top 16 and 12 sections of each core were analysed for organic matter, covering the period up to 1930. some of the layers from the core located in trois-rivières could not be analyzed since ephippial analysis consumed the available sediment amount. for sediment dating, core sections were freeze dried and analysed for gamma decay of radionuclides with an ortec hpge well detector (oak ridge, tn, usa). the resulting data were used to determine sediment dates and mass accumulation rates by applying the constant rate of supply (crs) model (appleby and oldfiel, 1983). radionuclides analysed included 210pb and 226ra, which together allowed the determination of unsupported 210pb (226ra minus 210pb) and supported 210pb (where 226ra and 210pb are in equilibrium) for the crs model. the artificially produced radionuclide 137cs, which has an expected peak in the study area in 1963 in association with nuclear weapons testing, was also analysed, in order to validate the 210pb chronology. image analysis of ephippia ephippia were manually isolated from sediments by examining the samples, with a dissecting scope. thereafter, ephippia were rinsed with demineralised water into a 100 µm sieve to eliminate residual particles. both the 1950’s and 1990’s layers for each of the four lakes were analysed for a total of eight samples, and an average of 20 ephippia were collected from each sample. each ephippium was digitized with a stereomicroscope nikon smz 745t joined to a ds-l3 camera unit under a standardized 50x zoom. even though pigmentation appeared symmetrical, the two sides of each ephippium were digitized in order to obtain an average of the data from both sides. microscope settings were standardized and each picture had the same preset threshold level for white balance, with constant lighting. biometric measures (length, width and total surface area of ephippia) were recorded with the dsl3 unit. digital pictures were thereafter processed with gimp 2.8.10 image manipulation software. each ephippium image was extracted from its background in order to analyze only the pixels from the ephippium. based on preliminary observations, two intensity colour thresholds have been selected for the pigmentation analysis (i.e., 75/255 and 125/255, fig. 1). these thresholds are in relation with a colour intensity scale of 255, in which 0 represents absolute black and 255 represents absolute white. the 75/255 tab. 1. main characteristics of the four study lakes, data measured in 1997 by the lmnpc (michel plante, lmnpc, unpublished results). lake average cladoceran chlorophyll a brook fish introduced doc area depth (ind l–1) (µg l–1) trout species species (mg l–1) (ha) (m) (n.) (n.) alphonse 13.2 5.5 0.76 0.95 present 2 1* 3.27 genévrier 4.0 8.0 0.72 1.17 extinct 0 0 4.66 giron 28.3 8.6 2.33 1.14 present 5 4** 4.69 noir 3.4 4.6 48.60 1.59 extinct 0 0 9.00 doc, dissolved organic carbon; *allegheny pearl dace (margariscus margarita); **allegheny pearl dace, common shiner (luxilus cornutus), northern redbelly dace (chrosomus eos) and brown bullhead (ameiurus nebulosus). whereas the latter is mainly benthivorous, all the other introduced species are potentially planktivorous. no n c om me rci al us e o nly 200 a. bérubé tellier et al. threshold enables quantification of the percentage of darkened pixels in the ephippium picture (located between 0 and 75 on the colour intensity scale of 255), while the 125/255 threshold provides the percentage of darkened and moderately darkened pixels in the picture (located between 0 and 125 on the colour intensity scale of 255). data registered represent the percentage of pixels from the selected image that are darker than the selected threshold (gerrish and caceres, 2003), and thus represent a proxy for the amount of melanin invested in the ephippium case. the selection of two different thresholds was motivated by the great range of variation in pigmentation among ephippia, where the variation in darker ephippia seemed better highlighted when using a threshold of 75 and that of the clearer ephippia by using a threshold of 125 (gerrish and caceres, 2003). we also used image analysis to extract measures of ephippia total area. ephippia were sorted into morphotypes (m1, m2 and m3, fig. 2) based on a visual examination of their characteristics, namely shape, the presence or absence of a spine, degree of symmetry, texture and length/width ratio. size and pigmentation were not used to classify ephippia, but instead used as dependent variables in the analyses. statistical analysis analyses were conducted only on the m1 morphotype to get a clearer picture of the observed variations. in order to take into account the nested nature of our sampling design (several ephippia per sample with two strata sampled in each lake), we used a mixed modelling approach to analyse our data. this approach not only allowed us to model properly the correlation among non-independent observations, but also to explicitly model heterogeneous variance if needed. to build our models, we applied the approach suggested by zuur et al. (2009) based on the comparison of the akaike information criterion (aic) among different models: i) we first selected the appropriate random term by comparing a full model fitted with all the independent fixed variables considered as important given the sampling design (lake type, stratum and their interaction), to an equivalent model with a random intercept for each study lake, and to another model with both a random intercept and slope for each study lake. restricted estimates maximum likelihood (reml) was used to calculate aic in this case; ii) we then selected the appropriate fixed terms by comparing the fit of the full model to a model without the interaction term and to a model without the stratum term (lake type only). maximum likelihood (ml) was used to calculate aic in this case; iii) once the random and the fixed terms were selected, the final model was refitted with reml to obtain a correct parameter estimation (zuur et al., 2009). all the models were fitted by using the lme()function in the nlme package in r. given that the preliminary exploration of the data suggested a problem of among-groups variance homogeneity, we included a heterogeneous variance term when needed by using the varident()function in the nlme package. three dependent variables were modelled with this approach: fig. 1. image analysis was used to find which percentage of the total pixels was darker than the threshold selected. a) original picture. b) picture processed with a threshold 75/255 and showing 34.5% dark pixels. c) picture processed with a threshold 125/255 and showing 64.4% dark pixels. no n c om me rci al us e o nly response of ephippium pigmentation to fish 201 ephippium surface (hereafter surface), percent ephippium pigmentation at threshold 75/255 (hereafter dark75) and 125/255 (hereafter dark125). all statistical analyses were performed in r 3.3.0 (r core team 2016). results the comparison of loi profiles for each lake strongly supports that duplicate cores have a similar stratigraphy (and also dates and mass sedimentation rates, supplementary fig. 1). for three lakes (alphonse, genévrier and noir), the accurate placement of the 137cs peak validated the 210pb dating. for lake giron, the peak for 137cs was later than expected (1990). dating the cores allowed the selection in each of the studied lakes of the 1950 and 1990 sediments layers, which represent the periods before and after park creation (supplementary fig. 2 and tab. 1). m1 resulted as the most ubiquitous morphotype (81% of analysed ephippia), being the only morphotype present in all of the study lakes, while m2 and m3 each accounted for 9% of the total ephippia and were found in one lake each. ephippium morphotype m1 was identified as belonging to species of the daphnia pulex leydig, 1860 group, based on two morphological identification keys created by vandekerkhove (2004) and mergeay et al. (2005). vandekerkhove’s key also permitted us to presume that the m3 morphotype is related to daphnia ambigua scoufield, 1946 species. however, the two identification keys did not provide enough information to identify the m2 morphotype. lakes with brook trout showed more diversity of morphotypes compared to fishless lakes, with lake alphonse containing both m1 and m2, and giron lake containing both m1 and m3. only the morphotype m1 was found in fishless lakes (genévrier and noir). since m1 resulted as the most common ephippium morphotype in all the studied lakes, statistical analyses were conducted only on the m1 morphotype to get a clearer picture of the observed variations. all selected models included a random intercept for the lake term and at least a term for heterogeneous variance: for both the surface and dark75 variables we included a term for heterogeneous variance across lakes, whereas for the dark125 variable, we also included a term for heterogeneous variance across strata. the visual representation by boxplots clearly illustrates this point (see the variables spread of boxplots among lakes in figs. 3 and 4). in all cases, the model selection for the fixed term ended up with a model including only the “lake type” factor, suggesting that neither the stratum, nor its interaction with the lake type were strong predictors for the three modelled variables. for the variable surface (fig. 3), the results of the t-test showed that lake type is not significantly related to dependent variable (p=0.174, tab. 2). in contrast, for both dark75 and dark125 variables (fig. 4), the results of the t-test showed that lake type is a significant predictor of the dependent variable (p=0.0238 and 0.0351 respectively, tab. 2), with darker ephippia found in fishless lakes and sediment layers. fig. 2. the three visually identified morphotypes (m1-m3). a) m1 (corresponding to d. pulex): spine, asymmetric shape, medium-sized margins, average length:width ratio of 1.42. b) m2: spine (broken on the picture), flare shape on both sides, narrow margin and average length/width ratio of 1.69. c) m3: no spine, spherical shape, large margins and average length:width ratio of 1.29; note the different scale for each picture (a 250 µm horizontal reference bar is presented on the bottom left of each specimen). no n c om me rci al us e o nly 202 a. bérubé tellier et al. discussion our study clearly indicated that ephippium pigmentation may be a better indicator of the presence of fish predation than ephippium size, a trait that showed a less clear pattern of variation between lakes with and without fish in our study system. as predicted, we found a sharp difference in pigmentation (both dark75 and dark125 variables) between ephippia collected from the 1990’s sediment layers from lakes with or without fish. ephippia collected from fishless lakes were significantly (both statistically, given the alpha level, and biologically, given the % variation between lake types) darker than ephippia from lakes with fish. although we did not directly analyze the optical properties of individual ephippia (nevalainen et al., 2016), the photographic approach used here (gerrish and cacéres, 2003) appeared to be sufficiently sensitive to track variations in ephippia pigmentation and to allow discriminating variations in fish predation pressure. small differences in size of ephippia between lake types were also detected, with ephippia showing a tendency to be larger in fishless lakes. unexpectedly, the same general pattern was also observed for the 1950’s period, suggesting that either brook trout extinction occurred before this period (e.g., at lake noir, for which no clear proof of brook trout presence is available for 1950), or that the population levels were already critically low before 1984 (e.g., at lake genévrier, when the last observation of brook trout is available), with resulting low predation pressure on zooplankton. the relatively small fig. 3. boxplot showing variations of ephippium surface according to lake and sediment layers (before and after park creation). lakes p1 (alphonse) and p2 (giron) have viable brook trout populations whereas lakes e3 (genévrier) and e4 (noir) were fishless in the 1990’s. grey: period before park creation; white: period after park creation. fig. 4. boxplots showing variations of the percentage of dark pixels at threshold 75 (panel a) and at threshold 125 (panel b) according to lake (viable or extinct) and sediment layer (before or after park creation). grey: period before park creation; white: period after park creation. no n c om me rci al us e o nly response of ephippium pigmentation to fish 203 size of the two fishless lakes could be related not only to a greater physical instability of these systems, but also to small fish population size, which could increase the vulnerability to stochastic events that could lead to extinction (dunham et al., 1999). it has already been shown that visual predators, such as brook trout, operate a strong selection pressure on large-sized cladocerans (brooks and dodson, 1965; galbraith, 1967), which in turn reduces of the average population body size (hart and bychek, 2011). predation pressure could eventually lead to a reduction in size at first reproduction, either by clonal replacement, or by phenotypic plasticity (latta et al., 2007), which could be mirrored by a reduction in average ephippium size. jeppesen et al. (2002), for example, observed a relationship between actual fish abundance and ephippia dorsal length in surface sediment. our data are in accordance with this hypothesis, with fishless lakes showing a tendency for larger ephippium size, compared to lakes with fish, which is related to adult body size. we also observed a high within-lake variability for this trait, which reduced the possibility to find significant differences given the low sample size of our study, and consequently, low power of our statistical analysis. nevertheless, the observation of clear differences in ephippium pigmentation between lake types highlights the sensitivity of this trait to variations in fish predation pressure. low levels of body pigmentation in the presence of visual predators have been observed in cladocerans (reinikainen, 2012), but to our knowledge this is the first study to show a similar phenomenon in ephippia. in fact, despite the great deal of variability in ephippium pigmentation among lakes (gerrish and cacéres, 2003), and the capability by visual predators of selectively removing individuals carrying darker ephippia (mellors, 1975), the relationship between population ephippium pigmentation and fish predation has not been established previously. since it has been shown that ephippium pigmentation is a strongly heritable trait (gerrish and cacéres, 2003), it is likely that the differences we observed between lake types are mainly due to clonal selection rather than plasticity per se. both selected measures of pigmentation (dark75 and dark125 variables) showed a significant relation to lake type. however, dark75 showed a clearer difference than dark125 in pigmentation variation between lake types (fig. 4). this finding suggests that lake type affected mainly the darkest range of pigmentation since dark75 was more selective than dark125 and integrated only the darkest pixels. this change might be related to an increased conspicuousness to visual predators when pigment concentration is higher (zaret and kerfoot, 1975). it was in fact observed that the quantity of pigmentation in the eyes of ceriodaphnia cornuta sars, 1885 is directly linked to predation risk, because its affects visibility (zaret, 1972). therefore, despite the high metabolic cost involved in melanin synthesis (hebert and emery, 1990), cladocerans produced darker ephippia in the absence of visual predators. this could be explained by the production of phenolic compounds associated to pigmentation, which can provide greater hardness and a better protection of the eggs to uv radiation, parasites and predation (zaret, 1972; hessen, 1996; gerrish, 2001; gerrish and caceres, 2003). in addition, ephippium pigmentation increases resistance to digestion by numerous planktivorous fish, fish-eating birds and mammals (mellors, 1975), which can provide cladocerans species a wider range of dispersal across a territory (proctor, 1964; proctor and malone 1965; mellors, 1975). therefore, it seems logical that in lakes where visual predators are absent, ephippium pigmentation is relatively high in order to increase hardness and enhance protection against uv radiation, parasites and predators. nevertheless, it is necessary to consider that pigmentation levels may be possibly driven by other factors affecting the uv risk in lakes, such as changes in dissolved organic carbon (doc) concentration (cooke et al., 2015) or in solar activity (nevalainen et al., 2016). in fact, variations in doc concentration, and in particular in its coloured or chromophoric component (cdom), can strongly control uv penetration in the water column, tab. 2. results for the fixed terms of the selected mixed linear models concerning the three studied variables. surface: ephippium total area expressed in pixels, dark75: percentage of dark pixels at threshold 75/255, dark125: percentage of dark pixels at threshold 125/255. variable value std. error df t-value p-value surface intercept 93,085.01 5783.30 124 16.10 <0.001 lake type (viable) -16,669.05 8038.32 2 -2.07 0.1738 dark75 intercept 66.98 3.80 124 17.61 <0.001 lake type (viable) -36.48 5.74 2 -6.36 0.0238 dark125 intercept 78.18 1.47 124 52.99 <0.001 lake type (viable) -12.15 2.34 2 -5.19 0.0351 df, degree of freedom. no n c om me rci al us e o nly 204 a. bérubé tellier et al. and can therefore potentially modulate pigmentation levels in cladocerans. similarly, variations in solar activity have been shown to be related to modulate uv-risk in fishless arctic ponds and, in turn, variations in ephippia melanization (nevalainen et al., 2016). however, in contrast to artic waters, which offer no strong protection against uv due to a lack of refugia, such as deep layers and shading macrophytes, our study lakes offer to zooplankton the possibility to avoid damaging uv radiation by changing their vertical or horizontal distribution during the day (williamson et al. 2011). thus, it is reasonable to suppose that ephippia pigmentation responded more strongly to fish predation than to variation in uvrisk in our systems. this might also explain why pigmentation levels tended to differ between our two fishless lakes, albeit the difference was clearly smaller than between fish and fishless lakes, with the brownwater lake noir (french for “black”) showing lower levels of ephippial pigmentation than lake genévrier, which has lower doc concentration. however, extrapolating actual doc concentration to the past decades has obvious limitations and more explicit analyses of the relationship between uv-risk and ephippial pigmentation along a gradient of doc concentration in natural lakes are necessary to elucidate this point. aic-based model selection did not support models including the time period or the interaction between the period and lake type, suggesting that ephippia produced before and after the creation of lmnpc have similar pigmentation levels in fishless lakes. however, a slight after vs before increase in pigmentation was observed at least for lake noir, suggesting a change in predation pressure. lake noir was mainly stocked with brook trout in the 1960’s, but there is no proof that populations were maintained in this lake until the park creation (plante, 1996b). even if they were present in the 1950’s, brook trout populations in this lake may not have been abundant enough to impose strong predation pressure on cladoceran populations. in that case, ephippia would not show the strong changes after the extinction of the remaining predators. the lack of time effect might also be related to physicochemical water status of those lakes, such as water turbidity (finlay et al., 2007), which can prevent efficient visual predation and thus hamper topdown effects on zooplankton (finlay et al., 2007). however, it is also important to note that our sampling design did not allow us to compare fishless lakes with lakes with brook trout only, because of the presence of one to three additional species of planktivorous fish in brook trout lakes. overall, this implies that the potential contrast between lake types is larger than between different time periods within lakes that experienced brook trout extinction. conclusions although this study did not show strong effects of lake types on average ephippium size, ephippium pigmentation was clearly correlated with the presence or absence of planktivorous fish in the 1990’s. our results also show that the level of planktivory in fishless lakes were very low in the 1950’s, suggesting either the absence of brook trout or very low population abundance. our study expanded the findings by jeppesen and co-workers (2002) by showing that the degree of ephippium pigmentation can be used in addition to ephippium size as a tool to quickly assess fish density in lakes when only sediment samples are available. complementary paleoecological analyses of other indicators of changes in the fish predation pressure (e.g., based on chaoborid mandibles, uutala 1990) are planned in order to strengthen these preliminary results on the impact of fish extinction in our study lakes. acknowledgments we would like to thank patricia bolduc, paméla magnan-baril, william brassard, freddy gutierrez trejo, chantal fournier, dany bouchard and joëlle guitard for their valuable help. we are greatly thankful to two anonymous reviewers for improving the first version of this work and to michel plante and his collaborators of lmnpc for their assistance with winter sampling and to parks canada for granting access to the study sites. references appelby pg, oldfield f, 1983. the assessement of 210pb data from sites with varying sediment accumulation rates. hydrobiologia 103:29-35. bertolo a, magnan p, plante m, 2008. liking the occurrence of brook trout with isolation and extinction in small boreal shield lakes. freshwater biol. 53:304-321. brooks jl, dodson si, 1965. predation, body size, and composition of plankton. science 150:28-35. cooke sl, fisher jm, kessler k, williamson ce, sanders rw, morris dp, porter ja, jeffrey wh, devaul princiotta s, pakulskim jd, 2015. direct and indirect effects of additions of chromophoric dissolved organic matter on zooplankton during large scale mesocom experiments in an oligotrophic lake. freshwater biol. 60:2362-2378. davidson ta, bennion h, jeppesen e, clarke gh, sayer cd, morley d, odgaard bv, rasmussen p, rawcliffe r, salgado j, simpson gl, amsinck sl, 2011. the role of cladocerans in tracking long-terme change in shallow lake trophic status. hydrobiologia 676:299-315. dunham j, peaock md, tracy cr, nielsen j, vinyard g, 1999. assessing extinction risk: integrating genetic information. conserv. ecol. 3:2. dzialowski ar, lennon jt, o’brien wj, smith vh, 2003. predno n c om me rci al us e o nly response of ephippium pigmentation to fish 205 ator-induced phenotypic plasticity in the exotic cladoceran daphnia lumholtzi. freshwater biol. 48:1593-1602. finlay k, beisner be, patoine a, pinel-alloul b, 2007. regional ecosystem variability drives the importance of bottom-up and top-down factors for zooplankton size spectra. can. j. fish. aquat. sci. 64:516-529. gerrish ga, 2001. maintaining pigment variation in aquatic systems: the cause and consequences of pigment variation in daphnia pulicaria ephippial egg casings. ms thesis, university of illinois at urbana-champaign, usa. gerrish ga, cacéres ce, 2003. genetic versus environmental influence on pigment variation in the ephippia of daphnia pulicaria. freshwater biol. 48:1971-1982. hansson la, 2004. plasticity in pigmentation induced by conflicting threats from predation and uv radiation. ecology 85:1005-1016. hebert pdn, emery cj, 1990. the adaptative significance of cuticular pigmentation in daphnia. funct. ecol. 4:703-710. hessen do, 1996. competitive trade-off strategies in arctic daphnia linked to melanism and uv-b stress. polar biol. 16:573-579. jeppesen e, jensen jp, amsinck s, landkildehus f, lauridsen t, mitchell sf, 2002. reconstructing the historical changes in daphnia mean size and planktivorous fish abundance in lakes from the size of daphnia ephippia in the sediment. j. paleolimnol. 27:133-143. lacasse s, magnan p, 1994. [distribution post-glaciaire de l’omble de fontaine dans le bassin hydrographique du fleuve saint-laurent: impact des interventions humaines]. [book in french]. université du québec à trois-rivières, pour le ministère de l’environnement et de la faune du québec, trois-rivières, qc, canada: 83 pp. latta l, bakelar jw, knapp ra, pfrender me, 2007. rapid evolution in response to introduced predators ii. bmc evol. biol. 7:21. masson s, pinel-alloul b, east p, magnan p, hogue g, barabé a, 2001. [programme de surveillance des écosystèmes aquatiques du parc national de la mauricie].[in french]. groupe de recherche en limnologie et environnement aquatique pour parcs canada, services de conservation des écosystèmes, parc national de la mauricie, qc. mellors wk, 1975. selective predation of ephippial daphnia and the resistance of ephippial eggs to digestion. ecology 56:974-980. mergeay j, verschuren d, de meester l, 2005. daphnia species diversity in kenya and a key to the identification of their ephippia. hydrobiologia 542:261-274. nevelainen l, rantala mv, luoto tp, ojala a, rautio m, 2016. long-term changes in pigmentation of arctic daphnia provide potential for reconstructing aquatic uv exposure. quaternary sci. rev. 144:44-50. plante m, 1996a. [plan de conservation des écosystèmes aquatiques].[in french]. parc national de la mauricie, parcs canada, service de conservation des ressources naturelles, parc national de la mauricie, qc. plante m, 1996b. [les communautés de poisson du parc national de la mauricie, de l’origine à aujourd’hui].[in french]. parcs canada, service de conservation des ressources naturelles, parc national de la mauricie, qc. proctor vw, 1964. viability of crustacean eggs recovered from ducks. ecology 45:656-658. proctor vw, malone cr, 1965. further evidence of the passive dispersal of small aquatic organisms via the intestinal tract of birds. ecology 46:728-729. r core team, 2016. r: a language and environment for statistical computing. r fundation for statistical computing. vienna, austria. available from: https://.r-project.org/ reinikanen m, ahlen e, 2012. ressurected ceriodaphnia quadrangula highlight differences between phenoand genotypic expressions. ecol. evol. 2:2989-2998. utala aj, 1990. chaoborus (diptera: chaoboridae) mandibulespaleolimnological indicators of the historical status of fish populations in acid-sensitive lakes. j. paleolimnol. 4:139-151. van der veen it, 2005. costly carotenoids: a trade-off between predation and infection risk? j. evol. biol. 18:992-999. vandekerkhove j, declerck s, vanhove m, brendonck l, jeppesen e, conde porcuna jm, de meester l, 2004. use of ephippial morphology to assess richness of anomopods: potentials and pitfall. j. limnol. 63:75-84. williamson ce, fisher jm, bollens sm, overholt ep, breckenridge jk, 2011. toward a more comprehensive theory of zooplankton diel vertical migration: integrating ultraviolet radiation and water transparency into the biotic paradigm. limnol. oceanogr. 56:1603-1623. zaret mt, 1972. invisible prey, and the nature of polymorphism in the cladocera (class crustacea). limnol. oceanogr. 17:171-184. zaret mt, kerfoot wc, 1975. fish predation on bosmina longirostris: body-size selection versus visibility selection. ecology 56:232-237. zuur af, leno en, walker nj, saveliev a, smith gm, 2009. mixed effects models and extensions in ecology with r. statistics. springer, new york: 574 pp. no n c om me rci al us e o nly layout 1 introduction the study of mucus secretion by cnidarians has seduced scientists for nearly a century (brown and bythell, 2005). mucus is produced to a greater or lesser extent by almost all cnidarians, in particular anthozoans and scyphozoans, and is essential for numerous life processes such as heterotrophic nutrition and sediment cleansing and as a defense against various environmental stressors and pathogens (parisi et al., 2020). much of the literature has examined the functions and composition of mucus in a limited number of cnidarians, investigating the trophic significance of mucus production, especially in coral reef species, but neglecting fundamental aspects of its ecological, physiological, and evolutionary role in a multitude of organisms belonging to this phylum (bythell and wild, 2011; myldlarz et al., 2016; bakshani et al., 2018; palmer and traylor-knowles, 2018). furthermore, the prevalence of issues related to climate change has redirected the research on implications related to this phenomenon, reducing the interest in the secretion of mucus in sensu stricto as a process underlying the adaptation and evolution of this ancient phylum. indeed, mucus is involved in several biological processes essential for almost all aquatic metazoan acting as the first contact stratum with surrounding media from cnidaria to chordata (stabili et al., 2015; alesci et al., 2022). the mucus layers provide a multifunctional and protective hydrogel interface between the organisms’ epithelial cells and the external environment (cabillon and lazado, 2019; kramar et al., 2019). aquatic invertebrates and fish use a layer of mucus to protect body surfaces, gills, and intestines. it effectively acts as a dynamic physical and biochemical barrier, showing numerous biological and ecological roles such as protection against abrasion, toxins, environmental pollutants and pathogens, and osmoregulation, chemical communication, parental nutrition and many others (hanaoka et al., 2001; ames et al., 2020). mucus has exceptional properties including elasticity, changing rheology and the ability to self-repair, thus representing an ideal means of trapping and immobilizing pathogens (brown and bythell, 2005). it is a key feature of the innate immune system in most aquatic and terrestrial metazoic phyla, playing a vital role in the prevention of microbial disease. in addition, while acting as a protective barrier, mucus allows the exchange of oxygen, carbon dioxide and nutrients and metabolites, lubricating the surfaces, reducing damage due to lesions, reducing dehydration of the epithelia and providing the polymer matrix which allows the transport of ciliary mucus particles (reverter et al., 2018; alesci et al., 2022). the properties of the secreted mucus generally depend on its composition consisting of exudates, such lipids, and mucins, polymeric glycoproteins that give the mucus its ability to gel, both released by mucocytes of review mucus secretions in cnidarian, an ecological, adaptive and evolutive tool serena savoca1, dario di fresco2, alessio alesci2, gioele capillo*3, nunziacarla spanò1 1department of biomedical, dental and morphological and functional imaging, university of messina, messina, italy; 2department of chemical, biological, pharmaceutical and environmental sciences, university of messina, messina, italy; 3department of veterinary sciences, university of messina, messina, italy abstract mucus secretion provides an interface with unique and multifunctional properties between the epithelial cells of many aquatic organisms and their surrounding environment. indeed, mucus is involved in various essential biological processes including feeding, reproduction, osmoregulation, competition for space, defense against pathogens, xenobiotics, and a multitude of environmental stressors. the ability to produce a functional mucus layer is an important evolutionary step, arising first in cnidaria that allowed for the development of the mucus-lined digestive cavity seen in higher metazoans. mucus secretion by cnidarians has been moderately investigated in both corals and jellyfish, which among cnidarians are the ones that have shown the highest secretion rates to date. however, although in corals the production of mucus has received more attention, especially in view of the important ecological role played in coral reefs, in medusozoans the topic is little considered. although the mucus secreted by corals has innumerable and important immunological, nutritional, and protective responsibilities, it should be remembered that jellyfish too represent a fundamental component of marine trophic web, playing numerous and important roles that are still unclear today. what is certain is that jellyfish are characterized (especially in the era of climate change) by large fluctuations in population density, the ecological implications of which are poorly understood. however, in both cases (medusozoans and anthozoans) to date some aspects relating to mucous secretions seem completely obscure, such as the microbiome and its variations as a function of environmental conditions or ontogenetic development, its implications in the field of immunological ecology, the consequent energy costs and finally the role played by the mucus in evolutionary terms. this review summarizes the properties, functions, ecological implications and evolutionary importance of mucus, in cnidarians, mainly focusing its roles in corals and jellyfish. understanding these aspects relating to the ecological and evolutionary importance played by mucus is of fundamental importance for the ecosystems functioning. no nco mm er cia l u se on ly s. savoca et al.72 the epithelium (brown and bythell, 2005). aging and some physical factors, such as depth and irradiance levels (crossland, 1987; zamzow, 2007), may influence the composition of the mucus, varying temporally within the same species and among different ones (crossland et al., 1980; meikle et al. 1988; ferrier-pages et al., 1998; hadaidi, 2019). it would seem that the main biophysical characteristics of mucus, like the ability to form gels of different viscosity and elasticity that lubricate and protect the underlying epithelia, depend on the type of mucins expressed (bythell and wild 2011). many hard and soft corals, for instance, continually release mucus, with a species-specific composition. mucus produced by cnidarians, similarly to other phyla, represents a fundamental aspect of the biology of these organisms and it is essential for several biological functions, including feeding, defending against pathogens, xenobiotics, and a multitude of environmental stressors, warding off aggression, and acting as an attacking weapon (baier et al., 1985; rivera-ortega and thomè, 2018; camachopacheco et al., 2022;). although continuous mucus releasing is physiological, several studies demonstrated that some environmental stressors, such as high particle concentrations or high sediment loads, and exposition to air due to low tide conditions, can increase mucus production (schuhmacher, 1977; rublee et al., 1980; krupp, 1984; wild et al., 2004; wild et al., 2005; liu et al., 2018;). additionally, there is limited information in the literature on mucus release as a consequence of ocean acidification and global warming (bythell and wild, 2011); however, during coral bleaching, caused by rising temperatures, an increase in mucus production has been detected (niggl et al., 2009; fransolet et al., 2012), possibly resulting in a mucus composition changes (wooldridge, 2009). anthozoan mucus has been studied for its biophysical characteristics, including rheology (activity under strain forces), and it has been discovered that it contains high molecular weight glycoproteins with characteristics that are similar to those of vertebrate mucins (jatkar et al., 2010), suggesting a similarity of mucus function between cnidarians and chordates (bythell & wild 2011). effectively, the ability to produce a functional mucosa represents an important evolutionary step, which first evolved in cnidaria and which was conserved into higher organisms, including terrestrial and human mammals. although cnidarians are widely used as model systems for evolutionary and biological investigations in metazoans, the importance of the evolution of a functional mucus layer in invertebrates is often overlooked (bakshani et al., 2018). another fundamental but little explored aspect is represented by the ecological role of mucus in the life strategy of cnidarians, especially in jellyfish (e.g., habitat for microbial colonization, involvement in biogeochemical cycles, dispersal of larvae, reproduction, and fertilization) (pitt et al., 2009; liu et al., 2018). zooxanthellae have been found to contribute substantially to the high gross primary production in coral reefs; however, corals release nearly half of the carbon assimilated by their zooxanthellae in the form of mucus (tarrant, 2007; bythell and wild, 2011;). the released mucus effectively traps organic matter from the water column and transports energy and nutrients to the sediment which acts as a biocatalytic mineralizing filter. in this way the mucus supplies energy to the heterotrophic reef community, thus establishing a recycling loop that supports benthic life, while reducing the loss of energy and nutrients from the reef ecosystem (wagner et al., 2012; goldberg, 2018). in this perspective, there is very little information on the ecological role of mucus in jellyfish. for example, up to 7% of the carbon assimilated by jellyfish is released into the environment in the form of mucus (hansson and norrman, 1995), which appears to play an important role in the carbon cycle (hansson and norrman, 1995; condon et al., 2011; tinta et al., 2021). however, jellyfish mucus has received little attention, resulting in a poor understanding of mucus excretion rates, composition, and its fate in the ecosystem. in the context of the expected increase in the jellyfish population, it is desirable to include jellyfish in energy flow models and ecosystem studies (ramondenc et al., 2020). lack of knowledge of the characteristics and functions of jellyfish mucus makes it more difficult to understand and model its role in the marine ecosystem. despite the multitude of ecological and physiological roles played by cnidarian mucus, and although some aspects of mucus secretion, function and composition have been widely investigated in some cnidarians, there are still many gaps in the knowledge of its composition variability, production rate, interand intraspecific variability, energetic significance, and ecological and evolutionary roles. the purpose of this review is therefore to examine what is known about cnidarian mucus in sensu stricto, sites and mechanisms of secretion, composition, function and ecological role and fate, providing an “evolutionary key reading”. cnidarian ecology: general features and evolution cnidarians occupy a key evolutionary position as basal metazoans and are ecologically important as predators, prey, and structure builders (goffredo and dubinsky, 2016; santos et al., 2016). the cnidarians are collectively an extremely diverse group of highly successful ancient aquatic organisms, which represent animals with incredibly different morphologies and life-history traits, but united by the presence of a specialized cell type: the cnidocyte (basso et al., 2019). molecular data suggest that the first cnidarians appeared about 740 million years ago (park et al., 2012). in 200 million years the phylum has undergone significant diversification (park et al., 2012) and today it consists of about 14,866 accepted species (worms, 2022), taxonomno nco mm er cia l u se on ly mucus secretions in cnidarian, an ecological, adaptive and evolutive tool 73 ically divided into five classes anthozoans, cubozoans, hydrozoans, scyphozoa and staurozoa, with an estimated number of species of ~11,000, including mainly marine invertebrates such as anemones, corals and jellyfish (daly et al., 2007). together, cnidarians have a global distribution and can occupy a wide range of both marine and freshwater environments. cnidarians present two forms, the polyp and the jellyfish, sessile and planktonic respectively. these are formed by three layers: an internal gastrodermis derived from the endoderm, surrounding the gastrovascular cavity, an intermediate matrix called mesoglea, and an external epidermis derived from ectoderm (lommel et al., 2018; d’ambra et al., 2021; zhao et al., 2021). the oral opening of the polyp is surrounded by tentacles, suitable for capturing the prey since heterotrophy is required to adequately supply nutrients for some species when photosymbiotic partners do not sufficiently supply the nutritional requirement (arai, 1997; goldberg, 2018). polyps can form colonies or live as solitary organisms. apart from the polyp stage, cubozoa, hydrozoa, scyphozoa, and staurozoa have a free-swimming planktonic stage, the jellyfish, which coexists with the polyp stage, but not all scyphozoa exhibit a polypoid phase, such as pelagia noctiluca. cubozoa, hydrozoa, and scyphozoa classes have a metagenic life cycle: this consists of a pelagic jellyfish that reproduces sexually and a benthic polyp that reproduces asexually (lucas et al., 2012). this suggests that they may have separated evolutionarily from anthozoa since most of them have only the polyp stage. cnidarians are phylogenetically basal aquatic animals belonging to metazoa, with radial symmetry and simple tissue organization. evolutionarily they represent the first metazoa; some of them can live for several centuries despite continuous exposure to different pathogens during their lifetime (petralia et al., 2014). since they possess only an innate immune system to protect themselves against pathogens, it is difficult to understand how cnidarians can achieve such high levels of longevity (parisi et al., 2020). phylogenetic analyses have confirmed that anthozoa appeared earlier than the other classes in the evolutionary history of the metazoa as they have circular dna, in contrast to cubozoa, hydrozoa, and scyphozoa, which have linear dna (technau and steele, 2011). due to their key position in evolution, these organisms play an essential ecological role in marine ecosystems and trophic web (tarrant, 2007). since cnidarians represent the most primitive metazoans, their evolution is of fundamental importance to study the origin of animal. the phylogeny of cnidaria has been debated for decades, and still remains unresolved for some taxa (kayal et al. 2018; mcfadden et al., 2022). the origin and evolution of cnidaria phylum posed the basis for their ecological success but the mechanisms lie at the origins and subsequent diversification of some characters, like symbiosis, colonial body plans, elaborate life histories, and cnidocytes, remain source of debate. however, a today accepted version of cnidaria phylogeny, based on morphology, life-history traits, and molecular data, is presented below. it is accepted that the common ancestor of all cnidaria is represented by a single polyp that lacked a medusoid stage. from this ancestor, around 15,100 species sorted in the following clades: anthozoa, endocnidozoa, and medusozoa. anthozoa comprise most of the species (~10,130) of cnidaria sorted in two accepted subclasses octocorallia and hexacorallia (see mcfadden et al., 2022 for a detailed phylogeny of anthozoa). endocnidozoa consist of around 798 species of endoparasites (myxozoa) and the monospecific polypodiozoa, polypodium hydriforme ussow, 1887 (a parasite of the oocytes of acipenseriform fishes (paddlefish and sturgeons). medusozoa encompass cubozoa (~48 species), hydrozoa (~3,875 species), scyphozoa (~241 species), and staurozoa (~49 species) (evans, et al. 2008; chang et al., 2015; zapata et al., 2015; kayal et al., 2017; mcfadden et al., 2022). the phylogenetic relationships of cnidaria are still discussed and in some cases far from being elucidated. despite this, recent studies using morphological, transcriptome, genome (both mitochondrial and nuclear), single gene, data tried to disentangle the complicated question (chang et al., 2015; zapata et al., 2015; kayal et al., 2017). despite the divergencies between the studies of chang es et al., 2015 and zapata et al., 2015, large congruence resulted in the evolutionary reconstructions of the cnidarian lineages, highlighting the importance of using large genomic datasets and increased taxon sampling in dealing evolutive questions (kayal et al., 2018). in this latter study, the authors, providing the most comprehensive taxon sampling of cnidaria for phylogenomic analysis, supported the monophyly of anthozoan, considering ceriantharia as the sister group of hexacoralia; and zoantharia as the sister group of all other hexacoralia. this, due to the presence in all the cited taxa of a single siphonoglyph (a ciliated groove located in the actinopharynx) shading light on the possible bilateral character of the anthozoa, and obviously of cnidaria, ancestor already proposed by finnerty et al., 2004. antipatharia were not included in the study by kayal et al., 2018, for the unavailability of genomic data for this group, that still represent a gap in evolutionary useful data for the cnidaria. for medusozoa the phylogenetic background is less entangled respect to anthozoa. in fact, it is accepted that scyphozoa, staurozoa, cubozoa, and hydrozoa, composing medusozoa, clade as two distinct monophyletic groups: i) the first, encompassing scyphozoa, staurozoa, and cubozoa, of which the earlier name is acraspeda in haeckel, 1880, that used the name with its present sense; and ii) the hydrozoa composed of monophyletic hydroidolina and trachylinae. finally, the sister relationship between endocnidozoa (myxozoa and polypodiozoa) and medusozoa, is corroborate by genomic no nco mm er cia l u se on ly s. savoca et al.74 analyses too (siddall et al., 1995; jimenez-guri et al., 2007; collins, 2009; chang et al., 2015; okamura et al., 2015; kayal et al., 2018; novosolov et al., 2022). depicting off an evolutionary processes conclusion it seems that all the cnidaria evolved from an ancestral nonsymbiotic, stinging, and medusa-lacking, solitary polyp with a planula larva (kayal et al., 2018; khalturin et al., 2019). symbiosis, coloniality, medusa life-history stages (also comprising losses of medusa stage events) occurred independently during the evolution of cnidaria. in addition to their key evolutionary position, cnidarians are also important as both predators and prey in aquatic ecosystems, playing a key role in biogeochemical cycles as recently suggested by tinta et al. (2021). gelatinous cnidarians are prey to several taxa including other cnidarians, fish and turtles. although gelatinous cnidarians are sometimes considered to be relatively unsightly or of low nutritional value (e.g., avent et al., 2001), in reality they are often underestimated as a food source because they tend to be digested quickly (arai et al., 2003; arai, 2005, costa et al., 2018). as predators, cnidarians can dramatically alter plankton composition and even compromise fishing yield through dietary overlap and direct predation (arai, 1988; purcell and arai, 2001; purcell and sturdevant, 2001). suspension-fed benthic cnidarians also impact planktonic composition and may play an important role in benthic-pelagic coupling. for example, a mediterranean gorgonian has been found to consume a large number of invertebrate eggs and larvae, in particular larval bivalves, with unknown effects on bivalve recruitment and consequently on population dynamics (rossi et al., 2004). particularly influential can be the aggregations or blooms of jellyfish (hydrozoa, scyphozoa or cubozoa) which can have important ecological consequences. recognition of the importance of jellyfish in marine ecosystems is growing steadily (tinta et al., 2021). the clade medusozoa includes four classes: cubozoa (carybdeida and chirodropida), hydrozoa (hydroidolina and trachylina), scyphozoa (coronatae and discomedusae), and staurozoa (cleistocarpida and eleutherocarpida). gelatinous cnidarians are found in all the world’s oceans and can occur in very high densities in large blooms; however, trends in jellyfish abundance appear to act on a local rather than a global scale (richardson et al., 2009; brotz et al., 2012; condon et al., 2013; schaub et al., 2018). for example, distinct coastal jellyfish hotspots that can span tens of square kilometres can occur in coastal waters (houghton et al., 2006). global jellyfish outbreaks appear to have become increasingly frequent and abundant in several coastal areas around the world and to last longer in recent years (heaslip et al., 2012). it is not yet clear whether this is just an upward phase of a natural pattern of decadal oscillations or a real increase in gelatinous zooplankton blooms (mccanch and mccanch, 1996; arai, 2005; purcell, 2005). it has been hypothesized that jellyfish have benefited from human-caused environmental changes, such as: climate change, overfishing, eutrophication, habitat modification (cardona et al., 2012; hamilton, 2016). outbreaks of jellyfish have caused concern about their potential harm to human welfare, tourism, and fisheries (graham et al., 2014). indeed, it would appear that due to their high abundance and spatial overlap, jellyfish could have a negative impact on fish of commercial interest through direct predation or through competition for plankton prey (decker et al., 2018). considering the extremely negative perception of jellyfish by the general public, it is not surprising that these organisms are often considered as stressors affecting the viability of ecosystem services (as in the case of competition with commercial fish stocks or reduction of the quality of bathing). in this regard, the role of jellyfish as “providers” of ecosystem services in terms of regulation, support, and provisioning should be redefined and highlighted. jellyfish have characteristics that place them in an influential position to structure the flow of energy through pelagic food webs, such as high growth and reproduction rates, large planktivorous diets, and few apparent predators (richardson et al., 2009). a wide range of taxa including other jellyfish, some molluscs, arthropods, fish, and birds regularly or opportunistically include jellyfish in their diet. even though jellyfish can have a low nutritional value compared to other prey items (doyle et al., 2007), consumption of sufficient quantities can sustain large predators. jellyfish are also extremely important predators in pelagic marine systems (pauly et al., 2009). the diversity in their feeding mechanisms and body size allows jellyfish to feed on a wide range of prey types and sizes (micro-heterotrophs, zooplankton, other jellyfish, fish, fish eggs and larvae). furthermore, most jellyfish are characterized by some sort of selectivity. when jellyfish blooms occur their rate of collective prey consumption can be so high as to directly or indirectly control the population size of other zooplanktonic organisms, including fish larvae (purcell, 1989). in particular, intense jellyfish predation on some prey items can cause a shift in the trophic structure of marine communities as a result of trophic cascades. thus, large jellyfish populations that can influence pelagic food webs, exerting top-down control over zooplankton communities (purcell, 2005; sommer and lengfellner, 2008; canepa et al., 2014; milisenda et al., 2018;). jellyfish can therefore significantly redirect the flow of energy through food webs and nutrient recycling (deason and smayda, 1982; suchman et al., 2008; condon et al., 2012; ruzicka et al., 2020). one of the most important services jellyfishes provide is climate regulation through the process of sequestering and transporting carbon through the water column. the sinking or accumulation of jellyfish carcasses on the seabed may also play an important role in the transfer of carbon from surface no nco mm er cia l u se on ly mucus secretions in cnidarian, an ecological, adaptive and evolutive tool 75 waters to the seabed (lebrato et al., 2012) through a complex pelagic-benthic coupling process. through the process of carbon sequestration, jellyfish provide regulating services (doyle et al., 2014), nutrients that support primary production (west et al., 2009), and organic matter that stimulates microbes. indeed, jellyfish can also provide small but significant amounts of nutrients to support primary production (pitt et al., 2009). inorganic nutrients such as carbon, nitrogen and phosphorus are “regenerated” by jellyfish excretion or also through mucus production. similarly, regenerated jellyfish products released into water are available to heterotrophic bacteria, and thus for respiration rather than primary production, creating a “jelly loop” involving the carbon cycle between jellyfish, bacteria, nanoflagellates heterotrophs, and ciliates (condon et al., 2011). numerous studies have investigated the role of bacteria during jellyfish bloom highlighting that variations in bacterial community structure associated with living or decaying jellyfish can have important consequences in trophic interactions and implications for carbon, nitrogen and phosphorus recycling (titelman et al., 2006; pitt et al., 2009; tinta et al., 2010; tinta et al., 2012; blanchet et al., 2015; tinta et al., 2016; apprill, 2017). bacteria associated with the outer surfaces of cnidarian epithelia may play a number of crucial roles, such as nitrogen fixation (lesser et al., 2004), antibiotic synthesis (ritchie and smith, 2004; kelman et al., 2006) decomposition of organic compounds (di salvo, 1969), primary defense against pathogens (ritchie, 2006). the surfaces of cnidarians, as well as other marine organisms, are a unique and favorable habitat for colonization by microorganisms and microbial communities often showing a marked variability (apprill, 2017). in fact, jellyfish can also provide important habitats for various species of marine organisms. the relationships and uses of jellyfish are so varied and complex that they resemble those typically described for tropical rainforests. jellyfish can provide (1) pelagic refuges or shelters, (2) pelagic substrate, and (3) a host for algal symbiotic associations. one of the best-documented biological interactions between jellyfish and marine organisms is the interaction between jellyfish and juvenile fish, generally considered a facultative symbiotic relationship (kramar et al., 2019). gelatinous cnidarians can provide a pelagic substrate (habitat) for a range of taxa from microbes to invertebrates, including crustaceans (perissinotto and pakhomov, 1998; pagès, 2000), pycnogonids (pagès et al., 2007), digeneans (martorelli, 2001) and protists (moss et al., 2001). studies on cnidarian jellyfish show the presence of endobiotic bacteria in tentacle jellyfish (kramar et al., 2019), as in the case of the upside-down jellyfish cassiopea sp. this spends most of the time upside down resting on the seabed in order to maximize the photosynthetic activity of the zooxanthellae present in their oral arms; their activity can increase benthic oxygen production nearly 100-fold (welsh et al., 2009). another algal host is the pleustonic jellyfish velella velella (purcell et al., 2012), in which its symbiotic zooxanthellae are able to photosynthesize and provide energy to the host. cnidarian-microrganisms interactions are well documented due to the extraordinary ecological importance of the mutualistic symbiosis between coralline anthozoans and cellular photosynthetic dinoflagellates (davy et al., 2012). cnidarians, mainly anthozoans, provide a three-dimensional structure to benthic ecosystems, particularly tropical coral reefs. coral reef ecosystems provide habitat for several taxa, protect coasts, and provide commercial and recreational resources for humans. indeed, several species of anthozoans play a vital role as bio-constructors, creating habitats that serve as shelter for many other organisms and are home to great biodiversity. anthozoans make up a class within this phylum and comprise a relatively well-known group consisting of hard and soft corals, gorgonians, sea feathers, black corals, and anemones (palmer and traylor-knowles, 2018). anthozoans are widely distributed throughout the seas of the world from the intertidal to the deep sea. unlike other classes of cnidarians, anthozoans never produce jellyfish during their life cycle (goffredo and dubinsky, 2016). within the anthozoans, the subclass octocorallia refers to the colonial species with eight tentacles arranged in a single cycle and has three orders: alcyonacea, helioporacea and pennatulacea. the other subclass, hexacorallia, includes species of anthozoans that can be solitary, colonial or aggregated, and consists of 6 orders: actiniaria, antipatharia, ceriantharia, corallimorpharia, zoanthinaria and scleractinia (otero et al., 2017). the latter are mainly responsible for the construction of coral reef structures, especially in tropical and subtropical seas, as well as cold-water corals (cwcs) that are among the main ecosystem engineering species in the deep sea worldwide and particularly in the mediterranean sea (soetaert et al., 2016; angiolillo et al., 2021). anthozoans are the most specious class of the phylum cnidaria, with an estimated 10,130 extant species (daly et al., 2007). anthozoans are phylogenetically basal, both within the metazoans as a whole and probably within cnidaria (kayal et al. 2013), with scleractinia already appearing in the middle triassic (about 250 million years ago (romano and palumbi, 1996). most existing anthozoans live in obligate endosymbiosis with dinoflagellates members of the genus symbiodinium (aranda et al., 2016), a relationship that underlies the ecological success of the class. in this association, symbiodinium provides over 90% of the energy needs of the anthozoan (muscatine and porter, 1977) and, for hard corals, facilitates the calcification of the exoskeleton, allowing the formation of coral reef ecosystems. in return, the anthozoan host protects its algal no nco mm er cia l u se on ly s. savoca et al.76 partner useful for the purpose of recycling waste carbon and nitrogen (jeong et al., 2012). this coral-scleractinian algae association is among the most investigated of the relationships that anthozoans have with the microbiota. this interaction could be closely related to the secretion of mucus in corals, with a preponderant role also in the ecology of the anthozoan immune system. cnidarians, particularly corals, possess a microbiota that is not only distinct from that of their surroundings, but also from other coral species (cooney et al., 2002; frias-lopez, 2002; ritchie and smith, 2004; brown and bythell, 2005). it is therefore plausible to hypothesize that the evolution of the superficial mucosa as a barrier may be associated with the ability to exclude certain bacterial species from body tissues, with the exception of a selected central microbiome. with a layer of mucus, as in cnidaria, non-commensal bacteria are essentially excluded from the tissues of the cnidarians. the activity of mucus in the exclusion of non-commensal bacteria may have initiated the evolution of the alimentary canal and therefore the evolution of higher organisms, as well as representing an essential feature of the innate immune system. it is therefore increasingly evident that the microbiome associated with anthozoans is crucial for their health and is possibly partially modulated by the immune mechanisms of corals (bourne et al., 2016). deciphering the immunological complexities of coral-microbe symbiosis is an ecologically key research field that could provide important insights into the establishment and functioning of symbiosis throughout the animal kingdom. this is particularly true when one considers that, despite their phylogenetic position and apparent morphological simplicity, anthozoans are immunologically complex (miller et al., 2007; shinzato et al., 2011), with large genomes and gene families that are comparable to those of bilateria (augustin and bosch, 2010), which is why they represent a very interesting group for the study of the evolution of immunity and mutualism. like all organisms, anthozoans possess innate immune mechanisms (palmer and traylor-knowles, 2012), but as invertebrates they lack the more complex adaptive component of immunity. innate immunity allows a non-specific and immediate response to both endogenous and exogenous threats with the aim of restoring homeostasis (beutler, 2004; medzhitov, 2008). however, at the same time, anthozoans utilize a large, complex, and diverse repertoire of immune receptors (miller et al., 2007), signaling pathways (wolenski et al., 2011) and “stress” responses (palmer et al., 2008), which allow to eliminate pathogens, heal wounds, and defend oneself (palmer and traylor-knowles, 2012; mydlarz et al., 2016). ecological immunology theory hypothesizes that variations between and within constituent immunity and immune responses are due to energy trade-offs between costly functional traits such as reproduction, growth, and maintenance / immunity (sheldon and verhulst, 1996; sadd and schmidhempel, 2009). in this regard, one of the first signs of suffering of corals is the increase in mucus secretion (brown and bythell 2005), a function that requires huge energy investments and that leads to the exhaustion of reserves, compromising the organism’s immunity (sheridan et al., 2014). similarly, thermal bleaching events result in the loss of symbiodinium which supplies energy to the coral causing it to die (brandt and mcmanus, 2009). the interruption of this interaction results in a state of energy deficit and malnutrition, which causes a lowering of the defenses and an increase in energy expenditure in order to resist or tolerate the disease (mayack and naug, 2009). site, mechanism of mucus secretion and composition in cnidarians mucus secretion is mainly attributed to specific cells, the mucocytes (goldberg, 2002; clarke et al., 2020), the only cells currently identified with mucus production in these organisms (brown and bythell 2005). however, other glandular cells have been identified in coral tissues (kinchington, 1981, le tissier, 1987; parker and krumlauf, 2017). mucocytes can occupy up to 90% of the ectoderm in some tissue areas: this has been confirmed by observation of histological sections in selected coral species (brown and bythell 2005). several studies highlighted the presence of mucocytes in three coral species: galexea fascicularis (linnaeus, 1767), tubastrea faulkneri (wells, 1982), acropora muricata (linnaeus, 1758) (marshall and wright, 1993). these three species belonging to class anthozoa, seem to be evolutionarily separate from the other classes of cnidarians, but they represent a good model for the description and mode of mucous secretion. in g. fascicularis, large mucocytes from 10 to 20 μm in size, were found on the external oral ectoderm, and in the oral gastrodermis. these cells were characteristically club-shaped, with basal nuclei and large granular inclusions. in the oral gastrodermis, they surrounded the zooxanthellae and presented granular inclusions larger than ectoderm mucocytes. in contrast, the aboral gastrodermis (opposite to the oral gastrodermis) was characterized by fewer and smaller mucocytes, 5 to 10 μm in length, although structurally comparable. few small mucocytes were also found in the calicoblastic layer, the portion of epidermis responsible for the formation of the skeletal structure. in t. faulkeri, mucocytes were very abundant in the calicoblastic layer and seem to be closely related to the skeletal structure. in a. muricata, no mucocytes were found in the apical portion of the oral ectoderm, although a thick layer of mucus covering the ectoderm was observed; in contrast, numerous mucocytes were found in the basal portion oral ectoderm (marshall & wright, 1993). this demonstrates that there is no single pattern of mucocytes distribution and abundance valid for all cnidarians, although such differences may depend on the physiological development/exchange no nco mm er cia l u se on ly mucus secretions in cnidarian, an ecological, adaptive and evolutive tool 77 of mucocytes (brown & bythell, 2005). it has been noted that due to induced stress the number of mucocytes increases. in a study by fransolet et al. (2013), the sea anemone exaiptasia diaphana (rapp, 1829) was subjected to a transient increase in temperature combined with high illumination for 30 hours; this resulted in bleaching, as endosymbiont microalgae belonging to the genus symbiondium disappeared. during the first day after the stress, a significant increase in cell proliferation was recorded at the gastrodermis and ectoderm levels, the ratio of ectodermal mucocytes increased significantly three weeks after the induced stress. the increase in mucocytes numbers may help restore symbiosis with endosymbiont dinoflagellates of the genus symbiondium (fransolet et al., 2013). in general, the main mucus functions are to lubricate the epithelium and keep it hydrated and provide protection from pathogens (mall, 2008). mucus prevents the passage of bacteria due to its particularly high viscosity while maintaining its impermeability to water and gases (cone, 2009). the ability of mucus to trap particles is closely related to its unique physicochemical properties as well as its phospholipid and glycolipid content (murty et al., 1984). an important characteristic of mucus is its layered structure: the layers are cyclically removed and regenerated, which allows contaminants to be efficiently eliminated, thus not reaching the underlying tissues (ducklow and mitchell, 1979), in a process known as sloughing, that further than playing an essential role in defense against pathogens, it determines the thickness of the mucus layer (cone, 2009). furthermore, mucus has elasticity, provided by the mucin fibres, which are composed of alternating glycosylated (hydrophilic) and bare (hydrophobic) protein regions; this conformation allows them to trap particles using a large number of low-affinity bonds that form and break very easily (cone, 2009). the mucus of the anemone actinia equina (linnaeus, 1758) contains biomolecules that exhibit lysozyme-like antibacterial activity, and it has been shown that the efficiency of these compounds is influenced by some environmental parameters, such as temperature, ph and ionic strength (stabili et al., 2015). however, these bioactive molecules have shown relevant antimicrobial effects under controlled conditions (temperature of 37 °c) that rarely occur in the natural environment of a. equina; this highlights the limitations of these compounds in performing their antibacterial activity at lower temperatures and more basic ph (stabili et al., 2015). similarly happens in hard and soft corals (class anthozoa) (kelman et al., 2006). however, the efficacy of antimicrobial compounds in the absence of mucus and the efficacy of mucus in the absence of antimicrobial compounds have not been tested (bakshani et al., 2018). in general, mucus in the animalia kingdom is predominantly composed of water (approximately 95% of its total wet mass), while the remaining 5% is composed of mucins, phospholipids, cholesterol, salt, lipids, fatty acids and proteins with defensive purpose, such as lysozyme, immunoglobulins and defensine (marshall and wright ,1991; celli et al., 2005; bansil & turner 2006; pearson et al., 2011;hubot et al., 2022). mucus of cnidaria is composed principally, as for other metazoa, of water at a percentage around 95%, with the remaining part composed by glycoproteins (�3%) and a mixture of other molecules (�2%), mostly represented by both inorganic salts (nacl) and organic compounds like peptides, nucleic acids, lipids, and antibodies (stabili et al., 2015; bakshani et al., 2018; hubot et al., 2022). these components, mainly glycoproteins, manage the mucus elasticity and viscosity (bansil and turner, 2018) useful in the defense (trough antibiotics and antimicrobial peptides), lubrification of underlying epithelia and act as a trap for food particles and transport to digestive system (bakshani et al., 2018; hubot et al., 2022). it has been reported the presence of nematocysts and toxins in the mucus of scyphomedusae (shanks and graham, 1988). mucus, when released, is inevitably contaminated by several elements, such as sediment, and a wide range of microrganisms: therefore, attempts have been made, in a controlled environment, to exploit artificial stresses such as dehydration and physical stimulation in order to achieve the release of mucus free of other components. however, any functional differences between mucus released under natural conditions and mucus released under artificial stress have not been investigated (brown and bythell, 2005). in this regard, previous work suggests that mucus obtained through artificial stress induction had a higher organic content than mucus taken in the natural environment (gottfried and roman, 1983). to our best knowledge, mucus macromolecular composition (proteins, lipids, and carbohydrates) and the c/n ratio in cnidarian has been described for few species including mainly belonging to anthozoa and scyphozoa (ducklow and mitchell, 1979; ducklow and mitchell, 1991; vacelet and thomassin, 1991; stabili et al., 2015; hubot et al., 2022). ducklow and mitchell (1991), studied the composition variability of the mucus of different anthozoa species, highlighting its utilization by microbial communities and the protective role of mucus especially related to water insolubility, chemical resistance to degradation, and inhibiting properties allowing a long degradation time by bacteria. in scyphozoa, hubot et al., 2022, evaluated the mucus macromolecules in the species aurelia aurita, chrysaora fulgida, chrysaora pacifica, eupilema inexpectata, and rhizostoma pulmo concluding a non-species-specific composition, highlighting the absence of data that demonstrate the variation in mucus biochemical features, related to different conditions and situations as feeding, reproduction and stress proposed by no nco mm er cia l u se on ly s. savoca et al.78 tinta et al., 2021. biological and ecological function of cnidarian mucus as stated above, most cnidarian species secrete mucus more or less constantly, which seems to be involved in both physiological and non-physiological functions. moreover, under certain stressful conditions, a considerable increase in mucus secretion rate can be observed (camachopacheco, 2022). from the literature is clear that the functions of mucus have been documented for some groups, especially in corals, where this seems to be closely related to their role as ecosystem engineers (bythell and wild, 2011). instead, the information about the role and the ecological and evolutionary implications of the mucus secreted by medusozoans is less substantial. basically, the secretion of mucus in invertebrates performs three essential functions, identified in defense, feeding and locomotion activities (brown and bythell, 2005). however, particularly in cnidarians, this assumes much more important roles and closely connected to the well-being of the cnidarians themselves and to the provision of ecosystem services. these include for example resistance to desiccation, sediment shedding, abrasion protection, defense against predators and pathogens, feeding, reproduction, settlement behaviour, spatial competition (stabili et al., 2015; camacho-pacheco, 2022). recently, some authors (liu et al., 2018; ames et al., 2020; hubot et al., 2022) have reported for some jellyfish species the mucus composition, excretion, and possible fate in marine ecosystems, which, considering the scarce information available in the literature, could represent an important piece in the understanding the biological and ecological function network claimed by mucosal products. several authors have eviscerated, demonstrated, or hypothesized all the possible biological functions of the mucus secreted by cnidarians, with particular attention paid to corals. however, our feeling is that few studies or reviews, currently present in the literature, have paid attention to the implications and ecological effects of the mucus produced by cnidarian (tinta et al., 2021). coral tissue is often heavily covered with cells that produce copious amounts of mucus (e.g., marshall and wright, 1993). although less well known, the epidermis and gastrodermis of jellyfish, including a. aurita, also contain numerous types of single-celled, mucus-producing glandular cells, leading to the formation of thin mucus layers that are constantly renewed on the outer surface of the jellyfish. under certain conditions such as stress, during reproduction, feeding and digestion, and even during death, the amount of mucus released is even more pronounced (rottini sandrini and avian, 1991; doyle et al., 2014). secretion of mucus sheets is known to be an important means of trapping and ingestion of small nutrient particles (e.g., yonge, 1930) by scleractinian corals and gorgonians (bessell-browne et al., 2017). lewis and price (1975) showed that increased mucus secretion by corals within different groups was a common response to the presence of food, which produced a dense net of filaments of various thicknesses, depending either on the species or water conditions. several jellyfish taxa have been listed as mucus net filter feeders (salps) (doyle et al., 2014). the production of mucus greatly facilitates the capture of prey by jellyfish due to the formation of lumps or conglomerates. studies of other jellyfish have reported that the mucus aids in the transport of food through the oesophagus to the gastric pouches, even when prey items were small (ames et al., 2020; daly et al., 2007). others can produce toxic mucus, such as cassiopea spp, which release large amounts of toxic mucus into the water column (ohdera et al., 2018; ames et al., 2020). cassiopeia mucus kills some fish species on contact (ames et al., 2020). motile cellular structures that trap and paralyze prey have been identified in the mucus of the jellyfish cassiopea xamachana (ames et al., 2020). in the order rhizostomae, mucus can facilitate feeding (arai, 1997), although its precise role and function are not yet fully understood (camacho-pacheco, 2022). the mucus is also involved in some reproductive processes, effectively becoming part of the reproductive or larval settlement strategy of some cnidarian species. a study conducted on the genus eudendrium (class hydrozoa) reported that polyps release planulae which remain attached to a thread of mucus which detaches once it reaches the bottom. this strategy allows the species to occupy welldefined sites, maintaining a gregarious behavior (wasserthal and wasserthal, 1973). a similar strategy has been found in nemertesia antennina (linnaeus, 1758, class hydrozoa), where the mucous filament reaches considerable lengths (up to 50 m) and allows the planulae to reach areas further away from the mother colony (hughes, 1975). in pelagia noctiluca, mucus increases the probability of fertility (lilley et al., 2014). first, each mature female medusa lays the oocytes in a mucus tape, holding the eggs together for several minutes before its dissolution. this peculiarity can favor the fertilization of the entire set of oocytes by the spermatozoa released by one or a few male partners (aglieri et al., 2014). only about 50% of the carbon and nitrogen invested during reproduction goes into the eggs, while the rest it produces a thread of mucus that binds all the eggs together. the mucus presumably increases the chances of fertilization and may increase the buoyancy of the egg series in pelagia (lilley et al., 2014). in a study performed on s. meleagris the greatest number of oocytes were found in the mucus, suggesting that it is also important in sexual reproduction, as previously reported for p. noctiluca (nagata and morandini, 2018) mucocyte secretions have distinct functions including cleansing sediment, healing wounds, and protecting against invasive microbes, as well as shielding from desiccation and uv damage (reviewed by brown and bythell 2005). increased mucus production is observed in no nco mm er cia l u se on ly mucus secretions in cnidarian, an ecological, adaptive and evolutive tool 79 colonies that are in direct contact with sunlight (remaining fully exposed during low tides) as a strategy to avoid desiccation (santos et al., 2016; sebens, 1982), as reported in the coral orbicella annularis (ellis & solander, 1786) (piggot et al., 2009). briefly, in sessile cnidarians the mucus represents a strong protection against desiccation (krupp, 1984), as its hydroscopic characteristics allow to keep the coral surface moist during exposure to air due to low tide conditions; however, the mechanisms that control these processes are still unclear. although the mucous layer is not waterproof, it can greatly reduce the airflow and water exchange rate. in anemones, shick (1991) observed that the mucus presents few barriers to the diffusion of water, but can produce a boundary layer which, when dried, separates from the external tissues (griffiths, 1977). another factor that may influence mucus secretion rates in corals is sedimentation (bak and elghershuizen, 1976; schumacher, 1977; rogers, 1990; stafford-smith and ormond 1992; stafford-smith, 1993). in some cases, repeated influxes of sediment can lead to depletion of mucus secretion causing the rejection of sediment to slow and cease (schumacher, 1977). in other species, such as gardinoseris planulata, on the other hand, mucus secretion continues long after the sediment rejection activity has ceased, suggesting that in this coral exhaustion of the ciliary mechanism rather than mucus secretion occurs (stafford-smith, 1993). stabili and collaborators (2015) demonstrated that some properties of the mucus secreted by a. equina, such as osmolarity and viscosity, as well as the chemical composition, play a key role in the defense against sedimentation and desiccation, however it is still evident that the hydrodynamism has a relevant influence on these phenomena. during an experimental manipulation, coffroth (1985) points out that the reduction of water movement also led to the formation of mucous sheets, thus the formation of the mucous sheet may be a secondary mechanism to cope with sediment stress, especially when currents are not strong enough to relieve environmental conditions. several studies demonstrated an increase in mucus secretion in corals due to sedimentation (bak and elghershuizen 1976; schumacher 1977; rogers, 1990; stafford-smith and ormond 1992; stafford-smith 1993; bongaerts et al. 2012; weber et al., 2012). this strategy turns out to be energetically quite expensive since (the carbon requirement for mucus secretion is more than doubled) (edmunds and davies,1989; riegl and branch, 1995). a study conducted on several species belonging to the orders scleractinia and malacalcyonacea showed that, in absence of sediment, mucus production was around 35% of daily respiration; in the presence of sediment, production increased up to 65%. therefore, sediments affect coral metabolise decreasing photosynthetic production, with greater carbon losses due to the greater amount of mucus produced (riegl and branch, 1995; weber et al., 2012). as a barrier defense, mucus on jellyfish surfaces has also been found to play a role in surface cleaning and predator defense (patwa et al., 2015). for scyphozoans, including s. meleagris, the mucus was found to contain nematocysts and toxins that serve defensive purposes (shanks and graham, 1988). shanks and graham (1988) identified mucus secretion as an important chemical defence mechanism against predators. in previous studies with aurelia coerulea, mucus was shown to function as a “cleaner”, preventing the accumulation of bacteria and debris on their surfaces (patwa et al., 2015). in addition to mechanical protection, the mucus of many cnidarians contains specific compounds to make the animal venomous, unpleasant, irritating or a combination of these characteristics. it is not surprising that these invertebrates have developed an innate immune system that produces a considerable number of defense molecules, such as lytic compounds (mayer et al., 2013), bioactive antimicrobials (aneiros and garateix, 2004; otero gonzalez et al., 2010; smith et al., 2010), toxins, and carbohydrate-based anti-adhesives (bavington et al., 2004). effectively the mucus acts as a physical barrier and the presence of cilia on the ectodermal cells favours the muco-ciliary transport of particles, including foreign cells, towards the mouth of the polyp where they can be digested or removed by the secretion of mucus. the first line of defense against invading pathogens in cnidarians is the superficial mucous layer. in reality, the superficial mucous layer represents the natural microbiota of the cnidarian surface, which may also include pathogens. the secreted mucus is an attractive niche for bacteria. since jellyfish mucus is composed primarily of proteins, lipids, and a lower proportion of carbohydrates (ducklow and mitchell, 1979), it is a high-quality energy source that is readily utilized by bacteria. cnidarians indeed host a variety of microbial species (rohwer et al., 2001; webster and bourne, 2007). a large number of evidence (reviewed by ainsworth et al., 2010) shows for example that “healthy” corals in their natural environment are associated with microbial groups and that when they are stressed or the environment changes, the associated microbes change in their turn. the cnidarian’s ability to control the production and composition of the superficial mucus layer and its associated bacteria may represent an important portion of their immunity. antibacterial compounds produced by the cnidarian host, as well as its associated microbial consortium, are released into the mucus and play an important role in the control of the mucus-associated microbial community (brown and bythell, 2005; reshef et al. 2006; ritchie 2006; rosenberg et al., 2007; shnitorland and kushmaro, 2009; rypien et al., 2010; krediet et al., 2013). the identification of the role of the microbiota in the development of the eukaryotic host and their response to environmental variations and/or stress, has led no nco mm er cia l u se on ly s. savoca et al.80 to the definition of a “holobiont”. there is now strong evidence indicating that the innate immune system of cnidarians is not only involved in the destruction of harmful microorganisms but is also crucial in the structuring of the microbial communities associated with the tissues that are essential components of the holobiont and with the health of the organism (rivera et al., 2015). cnidarians have many microorganisms associated with their tissues (rivera et al., 2015; tinta et al., 2019). actually, it is not entirely clear whether the bacteria adhere directly to the outer cell layers of the cnidarians or are associated only in the thin mucus layer, however, the association between bacteria and cnidarians is highly dynamic and complex. recent studies have focused on the relationships between microbes and their host organisms, investigating the composition and ecological role of microbial communities associated with cnidarians (basso et al., 2019; kramar et al., 2019; stabili et al., 2020). these studies addressed questions relating to the ecological role and composition of cnidarian-associated microbial communities, the mechanisms underlying these interactions, and the nature of the relationships that arise between cnidarians and their associated microbiome. further investigations involved bacteria associated with the outer surfaces epithelia in several cnidarian species and at different life stages, documenting their involvement in a multitude of important potential roles including antibiotic synthesis, nitrogen fixation, decomposition of organic compounds, the primary defence against pathogens or the modulation of contractile activities (weiland-bräuer et al., 2015; kramar et al., 2019; tinta et al., 2019). the microbial communities associated with the jellyfish species aurelia aurita, mastigias papua, cotylorhiza tuberculata and rhizostoma pulmo and tripedalia cystophora analyzed in several studies, showed extremely variable bacterial associations. the microbiome associated with different life stages of a. aurita (polyp, strobila, ephyra, juvenile and adult medusa) and with different compartments of adult specimens (exumbrella, mucus and gastric cavity), from different geographical areas was examined by weilandbräuer et al. (2015). the results of this study showed that the a. aurita microbial community appears to be highly host-specific and distinct from bacterioplankton suspended in the surrounding water column. in adults, the microbiota showed significant differences between the various compartments, showing a more variable bacterial composition associated with the mucus than that present in the gastric cavity. furthermore, kramar et al. (2019) monitored the bacterial community associated with the moon jellyfish aurelia solida. in particular, different parts of the body (umbrella surface, oral arms and gastric cavity) were analysed for bacterial community diversity. the authors reported that the microbiota associated with this species was different from the microbial assemblages in the surrounding seawater and differed significantly between different body compartments. furthermore, during the senescent phase, the bacterial community was found to mutate in structure with increased gammaproteobacteria (entirely vibrio). based on these results, it was hypothesized that the bacterial community associated with jellyfish may play an important role for the host. the compositional analysis of the microbiome associated with jellyfish mucus was conducted by tinta et al. (2019) to conclude that the gammaproteobacteria (mainly pseudoalteromonas and vibrio) are the most abundant, followed, to some extent also by the alphaproteobacteria (phaeobacter, rugeria and roseovarius). these bacteria are known for their ability to synthesize antimicrobial compounds when attached to live or inert surfaces, and therefore involved in defense of the host against pathogens and encrusting organisms from the surrounding environment. the microorganisms associated with the mucus of cnidarians appear to be characterized by a great diversity and the mucus represents the compartment richest in bacterial associates compared to the oral arms and the umbrella. in anthozoans, microbial agents perform key functions, including regulation of metabolism, immune defense, development, and behavior. bacteria associated with anthozoan tissues can fix nitrogen, digest complex polysaccharides, and produce antibiotics to prevent infection by pathogens. in turn, symbiodinium produces dimethylsulfoniopropionate (dmsp) as an osmolyte, an antioxidant agent (lawson et al., 2018) and a nutrient source for associated bacteria (rosenberg et al., 2009). a certain specificity has also been found in the anthozoa-bacteria association; porporato et al. (2013), for example, described the bacterial communities associated with pennatula phosphorea and pteroeides spinosum, showing a high species-specificity. moreover, since within the same species only a few phylotypes were shared between mucus and tissues, the authors also hypothesized that a partition of the microhabitat could exist between the associated microbial communities. in the case of p. phosphorea, both tissue and mucus associated communities were characterized by the predominance of alphaproteobacteria. conversely, alphaproteobacteria prevailed in the mucus layer of p. spinosum and the tissues were dominated by gammaproteobacteria. the isolates of strains belonging to vibrio spp., mainly obtained from coral mucus, showed an antibacterial activity against some indicator organisms, indicating a protective function of the bacterial communities associated with the coral as in the case of jellyfish. based on this evidence, the ability of cnidarians to control the production and composition of a mucous matrix and its associated bacteria may represent an important part of immunity (ocampo and cadavid, 2015). the ability of some bacteria to produce antimicrobial compounds is likely no nco mm er cia l u se on ly mucus secretions in cnidarian, an ecological, adaptive and evolutive tool 81 to contribute to competition for space and resources with potential pathogenic host bacteria. due to these characteristics, mucus and its components have interesting implications in molecular ecology and biotechnology. the specific interactions of microbial colonization of mucosal surfaces are still unknown. calow (1979) highlighted that any differences in the biochemical composition of the mucus can influence the attack of microbial agents that use exoenzymes to degrade the mucoid polymers. the microbes themselves can transform dissolved and particulate matter into living matter, attracting other predatory organisms. it also appears that changes in the composition and abundance of bacterial communities may affect the health of the host, making it more vulnerable to disease (reshef et al., 2006). recent documentation on the succession of microbial communities associated with the developmental stages of porites astreoides (sharp et al., 2012) and the discovery of beneficial functions in favor of alphaproteobacteria and marinobacter strains (lubbock, 1980) provide fundamental support to the hypothesis of “hologenome evolution”, i.e. the idea that in symbiotic organisms that were colonial, the hologenome, and thus sprung from all members of the holobiont, actually may act as a single unit of evolution, with rapidly evolving microorganisms providing the plasticity to adapt to the changing environment (rosenberg and zilber-rosenberg, 2011). the highly diverse mucosal microbiome is therefore generally believed to perform vital services and to be involved in the flow of energy within marine ecosystems. in this prospective, after detachment from the surface of cnidarians, mucus can play several important roles in the functioning of the ecosystem services it provides. first of all, the mucus acts as an energy vector (bythell and wild, 2011). some authors (haas et al., 2010; naumann et al., 2010; wild et al., 2010) confirm and demonstrate that all examined scleractinian corals release organic matter in the form of mucus, which includes both a particulate fraction and a dissolved fraction (mainly organic carbon, doc (crossland, 1987; naumann et al. 2010) and that mucus release by corals is largely independent of the carbon acquisition mechanism (autotrophy versus heterotrophy). studies conducted by naumann et al., 2010 suggest that photosynthates translocated by the endosymbiotic zooxanthellae of cnidarians represent the basis of mucus production (naumann et al. 2010). although the energy cost is quite high, (approximately 20-45% of the daily net photosynthate produced (brown and bythell, 2005), coral mucus is continuously released into the water column, accounting for half of the total mucus released by all benthic organisms (animals and plants) on coral reefs (crossland, 1987; naumann et al. 2010). suffice it to say that acropora spp., the dominant scleractinian genus on the great barrier reef, releases up to 4.8 l of mucus per square meter of coral reef surface per day (wild et al., 2004). zooxanthellae, endosymbiotic algae of reef-building corals, contribute substantially to the high gross primary production of coral reefs, supplying a substantial portion of their hosts’ energy needs by transferring photosynthetically fixed carbon to the coral. the high arabinose contents found in the carbohydrate fraction of coral mucus suggest that much of the fixed carbon is released in the form of mucus, since arabinose is not usually a constituent of animal cells (wild et al., 2004). between 56% and 80% of the released adhesive mucus dissolves in the reef water and effectively traps organic matter from the water column, increasing its initial organic carbon and nitrogen content by three orders of magnitude in short time and rapidly transporting energy and nutrients to the sediment, which acts as a biocatalytic mineralizing filter. coral mucus provides light energy transformed by zooxanthellae and trapped particles to the heterotrophic reef community, thus establishing a recycling cycle that supports benthic life, reducing energy and nutrient loss from the reef ecosystem (wild et al., 2004). however, mucus release can vary between species and depending on conditions. tanaka et al. (2009) found that only about 5% of the net daily photosynthetic production was released by acropora pulchra. muscatine et al. (1984) observed that in stylophora pistillata, depending on the irradiance levels, the loss of the newly fixed carbon varied from 6 to 50%. tremblay et al. (2012) found that in s. pistillata, the availability of heterotrophic foods and high light levels were required for the accumulation of autotrophic carbon, subsequently released as dissolved organic carbon. mucus release appears to constitute a dominant form of organic matter generated in coral reef ecosystems (e.g., ferrier-pagès et al., 1998; hatcher, 1988; bythell and wild, 2011) and represents the major route by which primary production of coral enters the food web (hatcher, 1988). freshly released coral mucus differs between species due to variations in lipid, sugar, and amino acid composition (ducklow and mitchell, 1979; crossland, 1987; meikle et al., 1988; wild et al. 2010). particulate mucus may also contain varying levels of phosphate or nitrogen, which may for example be influenced by planktonic food availability, colonization by picoplanktonic organisms and nitrogen-fixing bacteria (hubot et al., 2022). an old view held that coral mucus was a negligible source of nutrients (e.g., krupp, 1984; coffroth, 1990). today, however, it is known that mucus represents an energy substrate, rich in glucose (wild et al., 2010), degradable by microbes and higher organisms (benson and muscatine, 1974; grange, 1991; rinkevich et al., 1991; patton, 1994; naumann et al., 2010). several experimental studies (ducklow and no nco mm er cia l u se on ly s. savoca et al.82 mitchell, 1979; ferrier-pages et al., 2000; wild et al., 2004; 2005; 2009; 2010) have shown a stimulation of planktonic or benthic microbial activity after the addition of coral mucus. additional studies (allers et al., 2008; schöttner et al., 2009) have also shown that not only microbial activity, but also its diversity can be influenced by the presence of coral mucus. the cycles of production, aging and elimination of mucus are accompanied by variations in total microbial abundance, confirming that this colonization increases the nutritional value, in terms of carbon and nitrogen, of the released mucus (ferrier-pagès et al. 1998a; nakajima et al. 2009; grover et al. 2014; bednarz et al., 2017 and therein references). this enriched exudate becomes a downstream pelagic food source for fish (benson and muscatine, 1974) and zooplankton (gottfried and roman, 1983), and after sinking into the sediment, it is recycled back to members of the benthic community (wild et al., 2005; huettel et al., 2006; mayer and wild, 2010; tanaka et al. 2011; naumann et al., 2012). this results in a mechanism by which the pelagic food supply is coupled to the benthos (naumann et al., 2009; bythell and wild, 2011). in addition to its energy carrier function, the released adhesive mucus also functions, for obvious reasons, as a particle trap, which by forming aggregates of various inorganic and organic particles caught in the water column, greatly increases its sedimentation speed (wild et al., 2004; huettel et al., 2006). it is easy to hypothesize that such a mucus-induced and accumulated material cycle could support the recycling of essential nutrients within marine ecosystems, contributing to its functioning. a study done in the red sea documented that particle entrapment by coral mucus occurs while the mucus is still attached to the surface, due to the weaker velocities of tidal currents causing the mucus to remain longer at the surface of the coral (mayer and wild, 2010) than reported for great barrier reef, in which the mucus is rapidly detached (wild et al., 2004; huettel et al., 2006). this could result in a faster cycling of matter within the reef. thus, organic compounds can be rapidly recycled (wild et al., 2004; 2006), so that regenerated nutrients are released (wild et al., 2005) and feed new primary production by autotrophic coral reef organisms, including zooxanthellae. coral mucus can also trap picoplankton particles (davy and patten, 2007; futch et al., 2010) from the water column and transform them into larger mucus aggregates (naumann et al., 2009) allowing benthic filter of using picoplankton as a food source and once again facilitating the coupling between the water column and the benthic environment. even the mucus secreted by jellyfish can have important ecological implications, although less explored than corals. jellyfish can provide small but significant amounts of nutrients that support primary production (pitt et al., 2009). this is particularly important given the fact that jellyfish populations are characterized by large and rapid fluctuations in abundance. it is conceivable that the adaptive characteristics of jellyfish will allow them to thrive in anticipation of future catastrophic events directed by climate change, namely warming, acidification, oxygen loss and increasing human exploitation of ocean services; therefore, understanding the response of marine ecosystems to this natural and/or anthropic perturbation is of fundamental importance (richardson et al., 2009; purcell, 2012; steinberg and landry, 2017). regardless of the divergent scientific opinions on the causes of these fluctuations in jellyfish populations, the combined effect of natural oscillations and anthropogenic factors can lead to an increase in their populations, with consequent ecological and socio-economic impacts (richardson et al., 2009; purcell, 2012; condon et al. 2012, 2013; sanz-martín et al., 2016). precisely because of their demographic dynamics of boom and bust, jellyfish are likely to influence the cycles of matter in the ecosystems they inhabit. jellyfish blooms represent a significant and largely overlooked source of organic matter, especially at local and regional scales. recently, jellyfish have been recognized as key carbon export agents to the interior ocean, demonstrating the need to include jellyfish in oceanic biogeochemical models as a key component of the biological soft tissue pump (lebrato et al., 2012; steinberg and landry, 2017) so far not considered. jellyfish acquire c, n, and p by assimilating organic compounds from ingested prey, by absorbing small amounts of dissolved organic material, and some species instead actively assume dissolved inorganic forms. a portion of the ingested elements is incorporated into their biomass and the undigested material is eliminated in the faeces or released via “sloppy food”. the organic forms of c and n are recycled into the environment as mucus together with both organic and inorganic metabolic products. during bloom formation, populations increase in size acting as a net sink for c, n and p. during decomposition processes the elements bound to their biomass are “recycled” into the water column as dissolved inorganic and organic compounds, finally available for pelagic and benthic microbial communities. regenerated organic c by mucus production and decomposition supports microbial production, while regenerated inorganic n and p support algae production. additionally, few species, such as linuche unguiculata schwartz and many rhizostome species (genera cassiopea, mastigias and phyllorhiza) form a symbiosis with zooxanthellae. in zooxanthellate jellyfish, translocation of photosynthetic products from zooxanthellae is likely to be the major source of c for the host (cates, 1975). furthermore, inorganic excretory products can be translocated from the host to the zooxanthellae instead of being released into the external environment. once released into the environment, no nco mm er cia l u se on ly mucus secretions in cnidarian, an ecological, adaptive and evolutive tool 83 jellyfish mucus is metabolized by bacteria, creating important variations in microbial assemblages and diverting carbon to bacterial respiration (condon et al., 2011). mucus release therefore not only initiate biogeochemical cycles, but also control a range of processes including the microbial activity. so, again microbes are the main actor. as stated above, jellyfish host and interact with taxonomically and metabolically diverse microorganisms throughout their entire lives. the limited research available highlights the importance of the environmental microbial community for the recruitment of members of the jellyfish microbiome and a certain degree of microbiome specialization with some preferences for specific jellyfish taxa, life stages and body parts (basso et al., 2019; kramar et al., 2019; tinta et al., 2019). the role of the microbiota associated with jellyfish (or their mucus) is generally related to digestion, defense mechanisms against pathogenic microorganisms and possible predators (basso et al. 2019 and references therein), and reproduction (weiland-bräuer et al., 2020) collectively, the review of the current state of knowledge on the jellyfish microbiome reveals that this topic is grossly underestimated and should be studied in more detail in the future. mucociliary transport not only aids feeding by serving as a trap for nutrient particles, but also acts to remove nonnutritive particles from the surface of cnidarians. this is actually a common mechanism in several invertebrates, wherein the secreted mucus traps particulate debris which is then directed through the ciliated surface to a disposal site (sleigh, 1989). duerden (1906) was among the first to demonstrate the movement of mucus-bound non-nutritive particles up the edge of the disc of fungiid corals to the underlying substrate. it has been suggested that mucus have structural properties to efficiently accumulate, absorb or maybe bind pollutants, such as aromatic hydrocarbons (neff and anderson, 1981) or heavy metals (brown and howard, 1984; 1985), thus giving some protection to the underlying coral tissues by physically protecting them and acting as a pathway for pollutant release (neff and anderson, 1981). it seems likely that one of the primary detoxification mechanisms employed is increased mucus production. increased mucus secretion has been described in response to exposure to a wide range of pollutants: crude oil (mitchell and chet, 1975; neff & anderson, 1981), drilling mud (thompson et al., 1980), mercury (bastidas and garcia, 2004), copper sulphate (mitchell and chet, 1975), peat (dallmeyer et al., 1982). in histological studies, an increase in the number and size of mucosal secretory cells has been observed after exposure of manicina areolata (linnaeus, 1758) coral to chronic oil pollution (peters et al., 1981). in experiments using drilling mud, colonies of acropora cervicornis (lamarck, 1816) produced mucous filaments after 30 minutes of exposure, while mucus production by other species (porites astreoides lamarck, 1816, porites divaricata le sueur, 1820, porites furcata lamarck, 1816 and m. annularis) was observed only after 24 hours (thompson et al., 1980). an ecotoxicological study conducted by howe et al. (2012), tested copper toxicity on exaiptasia diaphana (rapp, 1829), a tropical symbiotic anemone; specifically, acute tests were conducted on juveniles (1-2 mm). the organisms responded to copper with severe (and often complete) retraction of tentacles, the collapse of the column, overproduction of mucus, the expulsion of grouped and individual zooxanthellae, and necrosis within 96 hours. these effects were rapid, with obvious changes in morphology occurring at concentrations ≥27 µg/l within the first hour. however, most anemones that survived acute copper exposure also survived continuous exposure, likely because they have successfully reduced their copper exposure and uptake through significant tissue retraction and employed detoxification mechanisms by increased mucus production, which allows the expulsion of metal-rich zooxanthellae. mucus produced by cnidarians can act as a trap for particles because of its adhesive properties. this can justify the presence of microplastic fibres found in cnidaria and ctenophora, as reported by a recent study conducted in orkney, shetland, and the north sea (devereux et al., 2021). this aspect deserves the attention of the scientific community, as the entry of microplastics through the cnidarians along the trophic chain can represent a potential risk both for ecosystems functioning and implications for human food safety and health due to the trophic transfer (barboza et al., 2018; de oliveira soares et al., 2020; albano et al., 2021; devereux et al., 2021; bruno et al., 2022). microplastic effects on freshwater cnidarians, particularly hydra thomseni (cordero, 1941) belonging to the class hydrozoa and on jellyfish of the genus aurelia were evaluated. these studies show that these organisms are capable of ingesting microplastics with associated impacts on feeding, causing significant morphological changes in h. thomseni. in aurelia sp., however, short-term exposure to microplastics compromises behavior as well as survival (murphy and quinn, 2018; costa et al. 2020). riegl and branch (1995) attempted to measure energy expenditure during coral cleaning processes, concluding that mucus secretion in sediment cleansing was energetically very costly. in fact, under stress, the carbon requirement for mucus secretion doubled. similar data were obtained by edmunds and davies (1989), who suggested that mucus secretion by porites porites stressed by the presence of non-food particles constituted an important pathway for energy loss. moreover, mucus overproduction has been also seen in cases of increased water temperature (neudecker, 1983) and decreased salinity (coffroth, 1985), associated with high energetic costs. climate change and anthropogenic stressors are threatening the long-term survival of many no nco mm er cia l u se on ly s. savoca et al.84 marine species, and coral reefs are among the most threatened ecosystems (hughes et al., 2017). long-lived organisms, such as scleractinian corals, are particularly vulnerable to environmental stresses and climate change, the rate of which occurs exceeds the time it takes for a population to adapt through natural selection (van oppen et al., 2017). one of the first signs of anthozoic suffering is represented precisely by the increased production of mucus (brown and bythell, 2005), which in turn requires huge energy investments, which can lead to the depletion of metabolic reserves, compromising the body’s immune system (riegl and branch, 1995; sheridan et al., 2014;). in acropora acuminata, energy expenditure through mucus exudation following environmental stress can account for up to 40% of the net carbon fixed by photosynthesis (crossland et al., 1980). an energy investment of this magnitude could therefore increase the risk of secondary adverse effects deriving from subsequent exposure to environmental stressors (anthony et al., 2009). subsequently, environmental stress could further deplete energy resources by inducing immune upregulation, known to be very expensive in terms of energy (armitage et al., 2003). unfavourable environmental conditions are compromising the health of many anthozoan species; global warming and therefore warmer and eutrophic waters, seem to promote the seizure of nutrient resources by symbiotic algae at the expense of their hosts (wooldridge, 2017; baker et al. 2018) and this shift from symbiosis to parasitism may presage the well-known expulsion of symbiotic coral partners and a change in the nature of their microbial communities (e.g., bourne et al., 2009; sokolow, 2009; work and meteyer, 2014; hughes et al. 2017). symbiotic relationships are highly sensitive to environmental change. regardless of their narrow temperature range, zooxanthellae become photosynthetically impaired, undergoing an expulsion process and varying degrees of bleaching. hosts in these conditions vary in their ability to obtain heterotrophic nutrition (grottoli et al., 2006; rodrigues and grottoli, 2007; palardy et al., 2008; ferrier-pagès et al. 2011), and if bleaching is severe, heterotrophy it must supply 100% of the metabolic needs of the coral (impossible for some species). thus, disruption of this mutualism results in a state of starvation and energy deficit, leaving the host vulnerable to infection, (e.g., sivajothy and thompson, 2002). similarly, infection is energetically very costly as resources are concentrated and invested in resisting or tolerating any disease (mayack and naug, 2009). according to ecological immunology theory, variations between and within constituent immunity and immune responses are due to energetic trade-offs between costly functional traits, such as reproduction, growth, and maintenance/immunity (sheldon and verhulst, 1996; sadd and schmid-hempel, 2009). therefore, studies of molecular ecology and immunity could offer a better understanding of the drivers that underlie the health status of both corals and other marine organisms in order to more effectively conserve and restore ecosystems of high ecological and socio-economic value such as the coral reefs. evolution of the mucus layer the development of the mucus layer is a key event in the evolutionary history of the kingdom animalia (bakshani et al., 2018). secretion of a functional mucus layer first appeared in the phyla cnidaria and ctenophora (lang et al., 2016). in a study published in 1984, the luxury carbon hypothesis was introduced, where the only function of mucus was to remove excess photosynthetic carbon from the coral-zooxanthellae symbiosis (davies, 1984). this hypothesis originated because mucus contains very high percentages of carbon. however, mucus did not evolve exclusively for this purpose, because this hypothesis is only applicable to cnidaria (bakshani et al., 2018). later, other important functions to the evolution of mucus were attributed, including capturing particles for feeding, preventing suffocation due to sediment, and providing a physical barrier that can keep pathogens out (edmunds and davies, 1989; riegl & branch, 1995; wild et al., 2004). sponges, belonging to the phylum porifera (which evolutionarily predates cnidaria), have genes that could be precursors of mucins (iwai et al., 2002; lang et al., 2007; 2016), so some sponges are even capable of producing mucus (biggerstaff et al., 2017; mcgrath et al., 2017). sponges are constantly in contact with environmental bacteria (found in abundance) (kennedy et al., 2008), and as cnidaria, have a distinct microbiota from the environmental one (cooney et al., 2002; frias-lopez et al., 2002; brown and bythell, 2005; savoca et al., 2019), with numerically fewer bacterial populations than porifera (ainsworth et al., 2010). the occurrence of microbes in mucus samples was confirmed by scanning electron microscopy (sem) and transmission electron microscopy (tem) microscopy analysis; however, they have not been found on the epithelial surfaces of polyps or on the exumbrella of jellyfish (the external part of the umbrella) (turk et al., 2021). cnidaria is often exploited as model organisms in developmental biology, yet the evolution of a functional mucus layer from invertebrates is rarely examined (bakshani et al., 2018), even though similarities are evident by comparing the cnidarians and vertebrates (including human) mucus composition (lai et al., 2009). bakshani et al. (2018), established the evolutionary conservation of the antimicrobial function of mucus by focusing on mucin genes and structure of mucus secreting cells among invertebrates and vertebrate. in the review, the authors resumed how secreted mucins evolved in early metazoans as cnidaria and the mechanisms lying at its bases is functional conserved from the first metazoans to no nco mm er cia l u se on ly mucus secretions in cnidarian, an ecological, adaptive and evolutive tool 85 mammals. conclusions and future research needs the importance of mucus in invertebrates, especially in cnidarians, is evident, as it is a fundamental secretion for the biology of these organisms and the marine ecosystems they inhabit. scleractinian corals have long been studied for their importance as ecosystem engineers and often used as model organisms for the study of symbiosis and reciprocal nutrient exchange. however, in examining the current state of knowledge on cnidarian mucus secretion, we found several gaps to be filled in the future. i) further efforts are required to determine and deepen the role played by the microbiome, its composition in species and the ecological implications of its speciesspecific variability, dependent on the life stage, or due to environmental stresses. similarly, it is not clear whether there is a substantial difference between organisms adhered to the surface and those present in the mucus and whether they act differently in the recycling of matter and/or in the flow of energy in coral ecosystems. ii) mucus production by scleractinian corals is prolific, but the amount of dissolved or particulate forms of this secretory material and their fate in the environment are still unresolved. iii) although the involvement of mucus in the innate immune system of cnidarians is clear, there are still discrepancies that prevent a complete evolutionary understanding of mucus. in this, and for many other aspects relating to the biological and ecological sphere of the cnidarians, studies on the hologenome could be particularly useful, involving all members of the holobiont and accepting the fact that this acts as a single unit of evolution, where the microorganisms, by adapting more rapidly, provide the plasticity necessary for the host to adapt to the rapidly changing, and probably for the worse, environment. iv) mucus secretion by jellyfish is a poorly explored field, although it has important ecological functions in antimicrobial defense, chemical defense and protection from environmental stress. the few existing studies on the jellyfish-microorganism association demonstrate the importance of these associations in various fundamental aspects for the life cycle, well-being and the ecological role played by these organisms. however, further studies on the interannual and seasonal dynamics of the associated jellyfish microbial community and possible variations under the pressure exerted by anthropogenic factors are needed. furthermore, new investigations are needed to produce potential common models of impacts of jellyfish biomass and their mucus production on both marine microbial communities and ecosystems worldwide. in particular, the link between jellyfish, mucus and microorganisms as final recipients and recyclers of the oceanic dissolved organic matter should be addressed more thoroughly. overall, an increased focus on this topic will contribute to understanding the dynamics of marine biodiversity and related ecological processes, as well as addressing the implications arising from climate change. v) it is somewhat demonstrated that the increase in mucous secretion in cnidarians has a high energy cost, which very often translates into an increased vulnerability of the organism. however, although described for a few coral species, it is largely misunderstood for cnidarians what the energy trade-off between costly functional traits (growth and reproduction) and maintenance/immunity is and how much energy investment is required for increasing mucus secretion in the anthropocene era. in conclusion, it is strongly suggested for these future studies that the combined use of eco-evo-devo approaches and ecological genomics would allow the individual disciplines and fields of investigation to enrich each other and would allow to improve not only the understanding of this topic in an evolutionary key, but also of the ecological, biological, and genetic mechanisms underlying the responses of organisms to their natural environments. corresponding author: gioele capillo, department of veterinary sciences, university of messina, polo universitario dell'annunziata, 98168, messina (italy). e-mail: gcapillo@unime.it authors’ contributions: all the authors made a substantive intellectual contribution. all the authors have read and approved the final version of the manuscript and agreed to be accountable for all aspects of the work. conflict of interest: the authors declare no potential conflict of interest. funding: none. availability of data and materials: all data generated or analyzed during this study are included in this published article. key words: cnidarian, mucus, ecological function, evolution, jellyfish, polyps. received: 30 november 2022. accepted: 16 december 2022. publisher’s note: all claims expressed in this article are solely those of the authors and do not necessarily represent those of their affiliated organizations, or those of the publisher, the editors and the reviewers. any product that may be evaluated in this article or claim that may be made by its manufacturer is not guaranteed or endorsed by the publisher. ©copyright: the author(s), 2022 licensee pagepress, italy advances in oceanography and limnology, 2022; 13:11054 doi: 10.4081/aiol.2022.11054 this work is licensed under a creative commons attribution noncommercial 4.0 international license (cc by-nc 4.0). no nco mm er cia l u se on ly s. savoca et al.86 references aglieri, g., papetti, c., zane, l., milisenda, g., boero, f., piraino, s., 2014. first evidence of inbreeding, relatedness and chaotic genetic patchiness in the holoplanktonic jellyfish pelagia noctiluca (scyphozoa, cnidaria). plos one. 9:e99647. ainsworth, t.d., thurber, r.v., gates, r.d., 2010. the future of coral reefs: a microbial perspective. trends ecol. evol. 25:233–40. albano, m., panarello, g., di paola, d., d’angelo, g., granata, a., savoca, s., capillo, g., 2021. the mauve stinger pelagia noctiluca (cnidaria, scyphozoa) plastics contamination, the strait of messina case. int. j. environ. stud. 78:977–82. alesci, a., pergolizzi, s., savoca, s., fumia, a., mangano, a., albano, m., et al., 2022. detecting intestinal goblet cells of the broadgilled hagfish eptatretus cirrhatus (forster, 1801): a confocal microscopy evaluation. biology (basel). 11:151–62. allers, e., niesner, c., wild, c., pernthaler, j., 2008. microbes enriched in seawater after addition of coral mucus. appl. environ. microbiol. 74:3274–8. ames, c.l., klompen, a.m.l., badhiwala, k., muffett, k., reft, a.j., kumar, m., et al., 2020. cassiosomes are stinging-cell structures in the mucus of the upside-down jellyfish cassiopea xamachana. commun. 3:67. aneiros, a., garateix, a., 2004. bioactive peptides from marine sources: pharmacological properties and isolation procedures. j. chromatogr. b anal. technol. biomed. life sci. 803:41– 53. angiolillo, m., la mesa, g., giusti, m., salvati, e., di lorenzo, b., rossi, l., et al., 2021. new records of scleractinian coldwater coral (cwc) assemblages in the southern tyrrhenian sea (western mediterranean sea): human impacts and conservation prospects. prog. oceanogr. 197:102656. anthony, k.r.n., hoogenboom, m.o., maynard, j.a., grottoli, a.g., middlebrook, r., 2009. energetics approach to predicting mortality risk from environmental stress: a case study of coral bleaching. funct. ecol. 23:539–50. apprill, a., 2017. marine animal microbiomes: toward understanding host–microbiome interactions in a changing ocean. front. mar. sci. 4:222. arai, m.n., 1988. interactions of fish and pelagic coelenterates. can. j. zool. 66:1913–27. arai, m.n., 1996. a functional biology of scyphozoa, a functional biology of scyphozoa. springer science & business media, amsterdam, the netherlands. 316 pp. arai, m.n., 2005. predation on pelagic coelenterates: a review. j. mar. biol. assoc. united kingdom. 85:523-36. arai, m.n., welch, d.w., dunsmuir, a.l., jacobs, m.c., ladouceur, a.r., 2003. digestion of pelagic ctenophora and cnidaria by fish. can. j. fish. aquat. sci. 60:825–9. aranda, m., li, y., liew, y.j., baumgarten, s., simakov, o., wilson, m.c., et al., 2016. genomes of coral dinoflagellate symbionts highlight evolutionary adaptations conducive to a symbiotic lifestyle. sci. rep. 6:1–15. armitage, s.a.o., thompson, j.j.w., rolff, j., siva-jothy, m.t., 2003. examining costs of induced and constitutive immune investment in tenebrio molitor. j. evol. biol. 16:1038–44. augustin, r., bosch, t.c.g., 2010. cnidarian immunity: a tale of two barriers. adv. exp. med. biol. 708:1–16. avent, s.r., bollens, s.m., butler, m., horgan, e., rountree, r., 2001. planktonic hydroids on georges bank: ingestion and selection by predatory fishes. deep. res. part ii top. stud. oceanogr. 48:673–84. baier, r.e., gucinski, h., meenaghan, m.a., wirth, j., glantz, p.q., 1985. biophysical studies of mucosal surfaces. oral interfacial react. bone, soft tissue saliva. bak, r.p.m., elgershuizen, j.h.b.w., 1976. patterns of oil-sediment rejection in corals. mar. biol. 37:105–13. baker, d.m., freeman, c.j., wong, j.c.y., fogel, m.l., knowlton, n., 2018. climate change promotes parasitism in a coral symbiosis. isme j. 12:921–30. bakshani, c.r., morales-garcia, a.l., althaus, m., wilcox, m.d., pearson, j.p., bythell, j.c., burgess, j.g., 2018. evolutionary conservation of the antimicrobial function of mucus: a first defence against infection. biofilms microbiomes 4:1–12. bansil, r., turner, b.s., 2006. mucin structure, aggregation, physiological functions and biomedical applications. curr. opin. colloid interface sci. 11:164–70. barboza, l.g.a., dick vethaak, a., lavorante, b.r.b.o., lundebye, a.k., guilhermino, l., 2018. marine microplastic debris: an emerging issue for food security, food safety and human health. mar. pollut. bull. 133:336-48. basso, l., rizzo, l., marzano, m., intranuovo, m., fosso, b., pesole, g., et al., 2019. jellyfish summer outbreaks as bacterial vectors and potential hazards for marine animals and humans health? the case of rhizostoma pulmo (scyphozoa, cnidaria). sci. total environ. 692:305–18. bastidas, c., garcia, e.m., 2004. sublethal effects of mercury and its distribution in the coral porites astreoides. mar. ecol. prog. ser. 267:133–43. bavington, c.d., lever, r., mulloy, b., grundy, m.m., page, c.p., richardson, n. v., mckenzie, j.d., 2004. anti-adhesive glycoproteins in echinoderm mucus secretions. comp. biochem. physiol. b biochem. mol. biol. 139:607–17. bednarz, v.n., grover, r., maguer, j.-f., fine, m., ferrier-pagès, c., 2017. the assimilation of diazotroph-derived nitrogen by scleractinian corals depends on their metabolic status. mbio 8:e02058-16. benson, a.a., muscatine, l., 1974. wax in coral mucus: energy transfer from corals to reef fishes 1. limnol. oceanogr. 19:810–4. bessel-browne, p., negri, a.p., fisher, r., clode, p.l., jones, r., 2017. impacts of light limitation on corals and crustose coralline algae. sci rep. 7:11553. beutler, b., 2004. inferences, questions and possibilities in tolllike receptor signalling. nature 430:257–63. biggerstaff, a., smith, d.j., jompa, j., bell, j.j., 2017. metabolic responses of a phototrophic sponge to sedimentation supports transitions to sponge-dominated reefs. sci. rep. 7:2725. blanchet, m., pringault, o., bouvy, m., catala, p., oriol, l., caparros, j., et al., 2015. changes in bacterial community metabolism and composition during the degradation of dissolved organic matter from the jellyfish aurelia aurita in a mediterranean coastal lagoon. environ. sci. pollut. res. 22:13638–53. bongaerts, p., hoeksema, b.w., hay, k.b., hoegh-guldberg, o., 2012. mushroom corals overcome live burial through pulsed inflation. coral reefs 31:399. bourne, d.g., garren, m., work, t.m., rosenberg, e., smith, no nco mm er cia l u se on ly mucus secretions in cnidarian, an ecological, adaptive and evolutive tool 87 g.w., harvell, c.d., 2009. microbial disease and the coral holobiont. trends microbiol. 17:554–62. bourne, d.g., morrow, k.m., webster, n.s., 2016. insights into the coral microbiome: underpinning the health and resilience of reef ecosystems. annu rev microbiol 70:317–40. branch, g.m., 1981. the biology of limpets: physical factors, energy flow, and ecological interactions. oceanogr. mar. biol. annu. rev. 19:235–380. brandt, m.e., mcmanus, j.w., 2009. dynamics and impact of the coral disease white plague: insights from a simulation model. dis. aquat. organ. 87:117–33. brotz, l., cheung, w.w.l., kleisner, k., pakhomov, e., pauly, d., 2012. increasing jellyfish populations: trends in large marine ecosystems, in: hydrobiologia. springer, amsterdam, the netherlands. p. 3–20. brown, b.e., bythell, j.c., 2005. perspectives on mucus secretion in reef corals. mar. ecol. prog. ser. 296:291–309. brown, b.e., howard, l.s., 1985. assessing the effects of “stress” on reef corals. adv. mar. biol. 22:1–63. bruno, c., blasi, m.f., mattei, d., martellone, l., brancaleone, e., savoca, s., favero, g., 2022. polymer composition analysis of plastic debris ingested by loggerhead turtles (caretta caretta) in southern tyrrhenian sea through atr-ftir spectroscopy. mar. environ. res. 179:105676. bythell, j.c., wild, c., 2011. biology and ecology of coral mucus release. j. exp. mar. bio. ecol. 408:88–93. cabillon, n.a.r., lazado, c.c., 2019. mucosal barrier functions of fish under changing environmental conditions. fishes. 4:2. calow, p., 1979. why some metazoan mucus secretions are more susceptible to microbial attack than others. am. nat. 114:149–52. camacho-pacheco, a.v., gómez-salinas, l.c., cisneros-mata, m.á., rodríguez-félix, d., díaz-tenorio, l.m., unzuetabustamante, m.l., 2022. feeding behavior, shrinking, and the role of mucus in the cannonball jellyfish stomolophus sp. 2 in captivity. diversity. 14:103. canepa, a., fuentes, v., sabatés, a., piraino, s., boero, f., gili, j.m., 2014. pelagia noctiluca in the mediterranean sea, in: jellyfish blooms. p. 237–66. cardona, l., álvarez de quevedo, i., borrell, a., aguilar, a., 2012. massive consumption of gelatinous plankton by mediterranean apex predators. plos one 7:e31329. cates, n., 1975. productivity and organic consumption in cassiopea and condylactus. j. exp. mar. bio. ecol. 18:55–9. celli, j., gregor, b., turner, b., afdhal, n.h., bansil, r., erramilli, s., 2005. viscoelastic properties and dynamics of porcine gastric mucin. biomacromolecules 6:1329–33. chang, e.s., neuhof, m., rubinstein, n.d., diamant, a., philippe, h., huchon, d., cartwright, p., 2015. genomic insights into the evolutionary origin of myxozoa within cnidaria. proc. natl. acad. sci. u. s. a. 112:14912–7. clarke, j.l., davey, p.a., aldred, n., aldred, n., 2020. sea anemones (exaiptasia pallida) use a secreted adhesive and complex pedal disc morphology for surface attachment. bmc zool. 5:1–13. coffroth, m.a., 1990. mucous sheet formation on poritid corals: an evaluation of coral mucus as a nutrient source on reefs. mar. biol. 105:39–49. coffroth, m.a., lasker, h.r., diamond, m.e., bruenn, j.a., bermingham, e., 1992. dna fingerprints of a gorgonian coral: a method for detecting clonal structure in a vegetative species. mar. biol. 114:317–25. collins, a.g., 2009. recent insights into cnidarian phylogeny. smithsonian. contributions to marine sciences. 38:139–49. condon, p., desbordes, g., miller, w.b., desteno, d., 2013. meditation increases compassionate responses to suffering. psychol. sci. 24, 2125–2127. condon, r.h., graham, w.m., duarte, c.m., pitt, k.a., lucas, c.h., haddock, s.h.d., et al., 2012. questioning the rise of gelatinous zooplankton in the world’s oceans. bioscience 62:160–9. condon, r.h., steinberg, d.k., del giorgio, p.a., bouvier, t.c., bronk, d.a., graham, w.m., ducklow, h.w., 2011. jellyfish blooms result in a major microbial respiratory sink of carbon in marine systems. proc. natl. acad. sci. u. s. a. 108:10225–30. cone, r.a., 2009. barrier properties of mucus. adv. drug deliv. rev. 61:75–85. cooney, r.p., pantos, o., le tissier, m.d.a., barer, m.r., o’donnell, a.g., bythell, j.c., 2002. characterization of the bacterial consortium associated with black band disease in coral using molecular microbiological techniques. environ. microbiol. 4:401–13. costa, e., gambardella, c., piazza, v., vassalli, m., sbrana, f., lavorano, s., et al., 2020. microplastics ingestion in the ephyra stage of aurelia sp. triggers acute and behavioral responses. ecotoxicol. environ. saf. 189:109983. costa, r., capillo, g., albergamo, a., volsi, r.l., bartolomeo, g., bua, g., et al., 2019. a multi-screening evaluation of the nutritional and nutraceutical potential of the mediterranean jellyfish pelagia noctiluca. mar. drugs 17:172. crossland, c.j., 1987. in situ release of mucus and doc-lipid from the corals acropora variabilis and stylophora pistillata in different light regimes. coral reefs 6:35–42. crossland, c.j., barnes, d.j., borowitzka, m.a., 1980. diurnal lipid and mucus production in the staghorn coral acropora acuminata. mar. biol. 60:81–90. d’ambra, i., merquiol, l., graham, w.m., costello, j.h., 2021. “indirect development” increases reproductive plasticity and contributes to the success of scyphozoan jellyfish in the oceans. sci. rep. 11:1–8. dallmeyer, d.g., porter, j.w., smith, g.j., 1982. effects of particulate peat on the behavior and physiology of the jamaican reef-building coral montastrea annularis. mar. biol. 68:229–33. daly, m., brugler, m.r., cartwright, p., collins, a.g., dawson, m.n., fautin, d.g., et al., 2007. the phylum cnidaria: a review of phylogenetic patterns and diversity 300 years after linnaeus. zootaxa. 1668:127–82. davies, j.m., viney, c., 1998. water-mucin phases: conditions for mucus liquid crystallinity. thermochim. acta. 315:39–49. davies, p.s., 1984. the role of zooxanthellae in the nutritional energy requirements of pocillopora eydouxi. coral reefs. 2:181–6. davy, s.k., allemand, d., weis, v.m., 2012. cell biology of cnidarian-dinoflagellate symbiosis. microbiol. mol. biol. rev. 76:229–61. davy, j.e., patten, n.l., 2007. morphological diversity of viruslike particles within the surface microlayer of scleractinian corals. aquat. microb. ecol. 47:37–44. no nco mm er cia l u se on ly s. savoca et al.88 de oliveira soares, m., matos, e., lucas, c., rizzo, l., allcock, l., rossi, s., 2020. microplastics in corals: an emergent threat. mar. pollut. bull. 161:111810. deason, e.e., smayda, t.j., 1982. ctenophore-zooplankton-phytoplankton interactions in narragansett bay, rhode island, usa, during 1972-1977. j. plankton res. 4:203–17. decker, m.b., robinson, k.l., dorji, s., cieciel, k.d., barceló, c., ruzicka, j.j., brodeur, r.d., 2018. jellyfish and forage fish spatial overlap on the eastern bering sea shelf during periods of high and low jellyfish biomass. mar. ecol. prog. ser. 591:57–69. devereux, r., hartl, m.g.j., bell, m., capper, a., 2021. the abundance of microplastics in cnidaria and ctenophora in the north sea. mar. pollut. bull. 173:112992. disalvo, l.h., 2021. 12 ingestion and assimilation of bacteria by two scleractinian coral species. exp. coelenterate biol. available from: https://www.degruyter.com/document/doi/10.1515/97808248 85335-014/pdf doyle, t.k., hays, g.c., harrod, c., houghton, j.d.r., 2014. ecological and societal benefits of jellyfish, in: pitt, k.a., lucas, c.h. (eds.), jellyfish blooms. springer dordrecht, netherlands. p. 105–127. doyle, t.k., houghton, j.d.r., mcdevitt, r., davenport, j., hays, g.c., 2007. the energy density of jellyfish: estimates from bomb-calorimetry and proximate-composition. j. exp. mar. bio. ecol. 343:239–52. ducklow, h.w., mitchell, r., 1979. bacterial populations and adaptations in the mucus layers on living corals 1. limnol. oceanogr. 24:715–25. ducklow, h.w., mitchell, r., 1991. composition of mucus released by coral reef coelenterates. limnol. oceanogr. 24:706–14. duerden, j.e., 1906. xxxiv.—the morphology of the madreporaria.—viii. the primary septa of the rugosa. j. nat. hist. 18:226–42. edmunds, p.j., davies, p.s., 1989. an energy budget for porites porites (scleractinia), growing in a stressed environment. coral reefs 8:37–43. evans, n.m., lindner, a., raikova, e. v., collins, a.g., cartwright, p., 2008. phylogenetic placement of the enigmatic parasite, polypodium hydriforme, within the phylum cnidaria. bmc evol. biol. 8:1–12. ferrier-pagès, c., peirano, a., abbate, m., cocito, s., negri, a., rottier, c., et al., 2011. summer autotrophy and winter heterotrophy in the temperate symbiotic coral cladocora caespitosa. limnol. oceanogr. 56:1429–38. ferrier-pagès, c., gattuso, j.p., cauwet, g., jaubert, j., allemand, d., 1998. release of dissolved organic carbon and nitrogen by the zooxanthellate coral galaxea fascicularis. mar. ecol. prog. ser. 172:265–74. finnerty, j.r., pang, k., burton, p., paulson, d., martindale, m.q., 2004. origins of bilateral symmetry: hox and dpp expression in a sea anemone. science. 304:1335–7. fransolet, d., herman, a.-c., roberty, s., plumier, j.-c., 2012. increased number of mucocytes in aiptasia pallida following bleaching. 12th international coral reef symposium. available from: https://www.icrs2012.com/proceedings/manuscripts/icrs2012_9a_5.pdf fransolet, d., roberty, s., herman, a.c., tonk, l., hoegh-guldberg, o., plumier, j.c., 2013. increased cell proliferation and mucocyte density in the sea anemone aiptasia pallida recovering from bleaching. plos one 8:e65015. frias-lopez, j., zerkle, a.l., bonheyo, g.t., fouke, b.w., 2002. partitioning of bacterial communities between seawater and healthy, black band diseased, and dead coral surfaces. appl. environ. microbiol. 68:2214–28. futch, j.c., griffin, d.w., lipp, e.k., 2010. human enteric viruses in groundwater indicate offshore transport of human sewage to coral reefs of the upper florida keys. environ. microbiol. 12:964–74. goffredo, s., dubinsky, z., (eds.), 2016. the cnidaria, past, present and future: the world of medusa and her sisters. springer, amsterdam, the netherlands. 855 pp. goldberg, w.m., 2002. feeding behavior, epidermal structure and mucus cytochemistry of the scleractinian mycetophyllia reesi, a coral without tentacles. tissue cell. 34:232–45. goldberg, w.m., 2018. coral food, feeding, nutrition, and secretion: a review. mar. org. as model syst. results and problems in cell differentiation. 65:377–421. gottfried, m., roman, m.r., 1983. ingestion and incorporation of coral-mucus detritus by reef zooplankton. mar. biol. 72:211–8. graham, w.m., gelcich, s., robinson, k.l., duarte, c.m., brotz, l., purcell, j.e., et al., 2014. linking human well-being and jellyfish: ecosystem services, impacts, and societal responses. front. ecol. environ. 12:515–23. grange, k.r., 1991. mutualism between the antipatharian antipathes fiordensis and the ophiuroid astrobrachion constrictum in new zealand fjords, in: hydrobiologia. springer, amsterdam, the netherlands. p. 297–303. griffiths, r.j., 1977. thermal stress and the biology of actinia equina l. (anthozoa). j. exp. mar. bio. ecol. 27:141–54. grottoli, a.g., rodrigues, l.j., palardy, j.e., 2006. heterotrophic plasticity and resilience in bleached corals. nature. 440:1186–9. grover, r., ferrier-pagès, c., maguer, j.-f., ezzat, l., fine, m., 2014. nitrogen fixation in the mucus of red sea corals. j. exp. biol. 217:3962–3. haas, a.f., naumann, m.s., struck, u., mayr, c., el-zibdah, m., wild, c., 2010. organic matter release by coral reef associated benthic algae in the northern red sea. j. exp. mar. bio. ecol. 389:53–60. hadaidi, g., gegner, h.m., ziegler, m., voolstra, c.r., 2019. carbohydrate composition of mucus from scleractinian corals from the central red sea. coral reefs 38:21–7. haeckel, e., 1880. essais de psychologie cellulaire. baillière, paris, france. 206 pp. hamilton, l.c., 2016. where is the north pole? an election-year survey on global change. available from: https:// scholars.unh.edu/carsey/285/ hanaoka, k., ohno, h., wada, n., ueno, s., goessler, w., kuehnelt, d., et al., 2001. occurrence of organo-arsenicals in jellyfishes and their mucus. chemosphere. 44:743–9. hansson, l.j., norrman, b., 1995. release of dissolved organic carbon (doc) by the scyphozoan jellyfish aurelia aurita and its potential influence on the production of planktic bacteria. mar. biol. 121:527–32. hatcher, b.g., 1988. coral reef primary productivity: a beggar’s banquet. trends ecol. evol. 3:106-11. no nco mm er cia l u se on ly mucus secretions in cnidarian, an ecological, adaptive and evolutive tool 89 heaslip, s.g., iverson, s.j., bowen, w.d., james, m.c., 2012. jellyfish support high energy intake of leatherback sea turtles (dermochelys coriacea): video evidence from animal-borne cameras. plos one. 7:e33259. houghton, j.d.r., doyle, t.k., wilson, m.w., davenport, j., hays, g.c., 2006. jellyfish aggregations and leatherback turtle foraging patterns in a temperate coastal environment. ecology. 87:1967–72. howe, p.l., reichelt-brushett, a.j., clark, m.w., 2012. aiptasia pulchella: a tropical cnidarian representative for laboratory ecotoxicological research. environ. toxicol. chem. 31:2653–62. hubot, n., giering, s.l.c., lucas, c.h., 2022. similarities between the biochemical composition of jellyfish body and mucus. j. plankton res. 44:337–44. huettel, m., wild, c., gonelli, s., 2006. mucus trap in coral reefs: formation and temporal evolution of particle aggregates caused by coral mucus. mar. ecol. prog. ser. 307:69–84. hughes, r.g., 1975. the distribution of epizoites on the hydroid nemertesia antennina (l.). j. mar. biol. assoc. 55:275–294. hughes, t.p., barnes, m.l., bellwood, d.r., cinner, j.e., cumming, g.s., jackson, j.b.c., et al., 2017. coral reefs in the anthropocene. nature. 546:82–90. iwai, t., inaba, n., naundorf, a., zhang, y., gotoh, m., iwasaki, h., et al., 2002. molecular cloning and characterization of a novel udp-glcnac: galnac-peptide β1,3-n-acetylglucosaminyltransferase (β3gn-t6), an enzyme synthesizing the core 3 structure of o-glycans. j. biol. chem. 277:12802–9. jatkar, a.a., brown, b.e., bythell, j.c., guppy, r., morris, n.j., pearson, j.p., 2010. coral mucus: the properties of its constituent mucins. biomacromolecules. 11:883–8. jeong, h.j., yoo, y. du, kang, n.s., lim, a.s., seong, k.a., lee, s.y., et al., 2012. heterotrophic feeding as a newly identified survival strategy of the dinoflagellate symbiodinium. proc. natl. acad. sci. u. s. a. 109:12604–9. jimnez-guri, e., okamura, b., holland, p.w.h., 2007. origin and evolution of a myxozoan worm. integr. comp. biol. 47:752–8. kayal, e., bentlage, b., pankey, m.s., ohdera, a.h., medina, m., plachetzki, d.c., et al., 2017. comprehensive philogenomic analyses resolve cnidarian relationships and the origins of key organismal traits. peerj preprints. 5:e3172v1 kayal, e., bentlage, b., pankey, m.s., ohdera, a.h., medina, m., plachetzki, d.c., et al., 2018. phylogenomics provides a robust topology of the major cnidarian lineages and insights on the origins of key organismal traits. bmc evol. biol. 18:1–18. kayal, e., roure, b., philippe, h., collins, a.g., lavrov, d. v, 2013. cnidarian phylogenetic relationships as revealed by mitogenomics. bmc evol. biol. 13:1–18. kelman, d., kashman, y., rosenberg, e., kushmaro, a., loya, y., 2006. antimicrobial activity of red sea corals. mar. biol. 149:357–363. kennedy, j., codling, c.e., jones, b. v, dobson, a.d.w., marchesi, j.r., 2008. diversity of microbes associated with the marine sponge, haliclona simulans, isolated from irish waters and identification of polyketide synthase genes from the sponge metagenome. environ. microbiol. 10:1888–902. khalturin, k., shinzato, c., khalturina, m., hamada, m., fujie, m., koyanagi, r., et al., 2019. medusozoan genomes inform the evolution of the jellyfish body plan. nat. ecol. evol. 3:811–22. kinchington, d., 1981. organic-matrix synthesis by scleractinian coral larval and post-larval stages during skeletogenesis, in: et al (eds) proc 4th int coral reef symp. pp. 107–113. kramar, m.k., tinta, t., lučić, d., malej, a., turk, v., 2019. bacteria associated with moon jellyfish during bloom and postbloom periods in the gulf of trieste (northern adriatic). plos one. 14:e0198056. krediet, c.j., ritchie, k.b., paul, v.j., teplitski, m., 2013. coralassociated micro-organisms and their roles in promoting coral health and thwarting diseases. proc. r. soc. b biol. sci. 280:20122328. krupp, d.a., 1984. mucus production by corals exposed during an extreme low tide. pacific sci. 38:1–11. lai, s.k., wang, y.y., wirtz, d., hanes, j., 2009. microand macrorheology of mucus. adv. drug deliv. rev. 61:86–100. lang, t., hansson, g.c., samuelsson, t., 2007. gel-forming mucins appeared early in metazoan evolution. proc. natl. acad. sci. u. s. a. 104:16209–14. lang, t., klasson, s., larsson, e., johansson, m.e.v., hansson, g.c., samuelsson, t., 2016. searching the evolutionary origin of epithelial mucus protein components mucins and fcgbp. mol. biol. evol. 33:1921–36. lawson, c.a., raina, j., kahlke, t., seymour, j.r., suggett, d.j., 2018. defining the core microbiome of the symbiotic dinoflagellate, symbiodinium. environ. microbiol. rep. 10:7–11. le tissier, m.d.a.a., 1991. the nature of the skeleton and skeletogenic tissues in the cnidaria, in: hydrobiologia. springer, amsterdam, the netherlands. p. 397–402. lebrato, m., pitt, k.a., sweetman, a.k., jones, d.o.b., cartes, j.e., oschlies, a., et al., 2012. jelly-falls historic and recent observations: a review to drive future research directions. hydrobiologia. 690:227–45. lesser, m.p., 2004. experimental biology of coral reef ecosystems. j. exp. mar. bio. ecol. 300:217–52. lewis, j.b., price, w.s., 1975. feeding mechanisms and feeding strategies of atlantic reef corals. j. zool. 176:527–44. lilley, m.k.s., elineau, a., ferraris, m., thiéry, a., stemmann, l., gorsky, g., lombard, f., 2014. individual shrinking to enhance population survival: quantifying the reproductive and metabolic expenditures of a starving jellyfish, pelagia noctiluca. j. plankton res. 36:1585–97. liu, w., mo, f., jiang, g., liang, h., ma, c., li, t., et al., 2018. stress-induced mucus secretion and its composition by a combination of proteomics and metabolomics of the jellyfish aurelia coerulea. mar. drugs. 16:341. lommel, m., strompen, j., hellewell, a.l., balasubramanian, g.p., christofidou, e.d., thomson, a.r., et al., 2018. hydra mesoglea proteome identifies thrombospondin as a conserved component active in head organizer restriction. sci. rep. 8:1–18. lubbock, r., 1980. clone-specific cellular recognition in a sea anemone. proc. natl. acad. sci. 77:6667–9. lucas, c.h., graham, w.m., widmer, c., 2012. jellyfish life histories: role of polyps in forming and maintaining scyphomedusa populations. adv. mar. biol. 63:133–96. mall, a.s., 2008. analysis of mucins: role in laboratory diagnosis. j. clin. pathol. 61:1018–24. marshall, a.t., wright, o.p., 1993. confocal laser scanning light no nco mm er cia l u se on ly s. savoca et al.90 microscopy of the extra-thecal epithelia of undecalcified scleractinian corals. cell tissue res. 272:533–43. marshall, a.t., wright, o.p., 1991. freeze-substitution of scleractinian coral for confocal scanning laser microscopy and xray microanalysis. j. microsc. 162:341–54. martorelli, s.r., 2001. digenea parasites of jellyfish and ctenophores of the southern atlantic. hydrobiologia. 451:305–10. mayack, c., naug, d., 2009. energetic stress in the honeybee apis mellifera from nosema ceranae infection. j. invertebr. pathol. 100:185–8. mayer, a.m.s., rodríguez, a.d., taglialatela-scafati, o., fusetani, n., 2013. marine pharmacology in 2009–2011: marine compounds with antibacterial, antidiabetic, antifungal, antiinflammatory, antiprotozoal, antituberculosis, and antiviral activities; affecting the immune and nervous systems, and other miscellaneous mechanisms of action. mar. drugs. 11:2510–73. mayer, f.w., wild, c., 2010. coral mucus release and following particle trapping contribute to rapid nutrient recycling in a northern red sea fringing reef. mar. freshw. res. 61:1006–14. mccanch, n. v, mccanch, m., bell, j., 1996. fulmars feeding on jellyfish. br. birds 89:569. mcfadden, c.s., van ofwegen, l.p., quattrini, a.m., 2022. revisionary systematics of octocorallia (cnidaria: anthozoa) guided by phylogenomics. available from: https://ssbbulletin.org/index.php/bssb/article/view/8735 mcgrath, e.c., smith, d.j., jompa, j., bell, j.j., 2017. adaptive mechanisms and physiological effects of suspended and settled sediment on barrel sponges. j. exp. mar. bio. ecol. 496:74–83. medzhitov, r., 2008. origin and physiological roles of inflammation. nature. 454:428–35. meikle, p., richards, g.n., yellowlees, d., 1988. structural investigations on the mucus from six species of coral. mar. biol. 99:187–93. milisenda, g., rossi, s., vizzini, s., fuentes, v.l., purcell, j.e., tilves, u., piraino, s., 2018. seasonal variability of diet and trophic level of the gelatinous predator pelagia noctiluca (scyphozoa). sci. rep. 8:1–13. miller, d.j., hemmrich, g., ball, e.e., hayward, d.c., khalturin, k., funayama, n., et al., 2007. the innate immune repertoire in cnidaria ancestral complexity and stochastic gene loss. genome biol. 8:1–13. mitchell, r., chet, i., 1975. bacterial attack of corals in polluted seawater. microb. ecol. 2:227–33. moss, a.g., estes, a.m., muellner, l.a., morgan, d.d., 2001. protistan epibionts of the ctenophore mnemiopsis mccradyi mayer, in: hydrobiologia. 451:295–304. murphy, f., quinn, b., 2018. the effects of microplastic on freshwater hydra attenuata feeding, morphology & reproduction. environ. pollut. 234:487–94. murty, v.l.n., sarosiek, j., slomiany, a., slomiany, b.l., 1984. effect of lipids and proteins on the viscosity of gastric mucus glycoprotein. biochem. biophys. res. commun. 121:521–9. muscatine, l., falkowski, p.g., porter, j.w., dubinsky, z., 1984. fate of photosynthetic fixed carbon in light-and shadeadapted colonies of the symbiotic coral stylophora pistillata. proc. r. soc. london. ser. b. biol. sci. 222:181–202. muscatine, l., porter, j.w., 1977. reef corals: mutualistic symbioses adapted to nutrient-poor environments. bioscience. 27:454–60. mydlarz, l.d., fuess, l., mann, w., pinzón, j.h., gochfeld, d.j., 2016. cnidarian immunity: from genomes to phenomes, in: goffredo, s., dubinsky, z. (eds.), the cnidaria, past, present and future: the world of medusa and her sisters. springer international publishing, cham, switzerland. p. 441–66. nagata, r.m., morandini, a.c., 2018. diet, prey selection, and individual feeding rates of the jellyfish lychnorhiza lucerna (scyphozoa, rhizostomeae). mar. biol. 165:187. nakajima, r., yoshida, t., azman, b.a.r., zaleha, k., othman, b.h.r., toda, t., 2009. in situ release of coral mucus by acropora and its influence on the heterotrophic bacteria. aquat. ecol. 43:815–23. naumann, m.s., mayr, c., struck, u., wild, c., 2010. coral mucus stable isotope composition and labeling: experimental evidence for mucus uptake by epizoic acoelomorph worms. mar. biol. 157:2521–31. naumann, m.s., richter, c., el-zibdah, m., wild, c., 2009. coral mucus as an efficient trap for picoplanktonic cyanobacteria: implications for pelagic–benthic coupling in the reef ecosystem. mar. ecol. prog. ser. 385:65–76. naumann, m.s., richter, c., mott, c., el-zibdah, m., manasrah, r., wild, c., 2012. budget of coral-derived organic carbon in a fringing coral reef of the gulf of aqaba, red sea. j. mar. syst. 105:20–9. neff, j.m., anderson, j.w., 1981. response of marine animals to petroleum and specific petroleum hydrocarbons. j of appl. ecol. 19:674. neudecker, s., 1981. growth and survival of scleractinian corals exposed to thermal effluents at guam, in: proceedings of the 4th international coral reef symposium. available from: http://www.reefbase.org/pacific/pub_a0000000049.aspx niggl, w., glas, m., laforsch, c., mayr, c., wild, c., 2009. first evidence of coral bleaching stimulating organic matter release by reef corals. proceedings of the 11th international coral reef symposium. available from: https://nsuworks. nova.edu/cgi/viewcontent.cgi?filename=185&article=1000& context=occ_icrs&type=additional novosolov, m., yahalomi, d., chang, e.s., fiala, i., cartwright, p., huchon, d., 2022. the phylogenetic position of the enigmatic, polypodium hydriforme (cnidaria, polypodiozoa): insights from mitochondrial genomes. genome biol. evol. 14:evac112. ocampo, i.d., cadavid, l.f., 2015. mechanisms of immune responses in cnidarians. acta biológica colomb. 20:5–11. ohdera, a.h., abrams, m.j., ames, c.l., baker, d.m., suescúnbolívar, l.p., collins, a.g., et al., 2018. upside-down but headed in the right direction: review of the highly versatile cassiopea xamachana system. front. ecol. evol. 35. okamura, b., gruhl, a., bartholomew, j.l., 2015. an introduction to myxozoan evolution, ecology and development, in: myxozoan evolution, ecology and development. springer, amsterdam, the netherlands. p. 1–20. otero, m. del m., numa, c., bo, m., orejas, c., garrabou, j., cerrano, c., et al., 2017. overview of the conservation status of mediterranean anthozoa. available from: https://portals.iucn.org/library/node/46713 otero-gonzáiez, a.j., magalhães, b.s., garcia-villarino, m., no nco mm er cia l u se on ly mucus secretions in cnidarian, an ecological, adaptive and evolutive tool 91 lópez-abarrategui, c., sousa, d.a., dias, s.c., franco, o.l., 2010. antimicrobial peptides from marine invertebrates as a new frontier for microbial infection control. faseb j. 24:1320–34. pages, f., 2000. biological associations between barnacles and jellyfish with emphasis on the ectoparasitism of alepas pacifica (lepadomorpha) on diplulmaris malayensis (scyphozoa). j. nat. hist. 34:2045–56. pagès, f., corbera, j., lindsay, d., 2007. piggybacking pycnogonids and parasitic narcomedusae on pandea rubra (anthomedusae, pandeidae). plankt. benthos res. 2:83–90. palardy, j.e., rodrigues, l.j., grottoli, a.g., 2008. the importance of zooplankton to the daily metabolic carbon requirements of healthy and bleached corals at two depths. j. exp. mar. bio. ecol. 367:180–8. palmer, c. v., traylor-knowles, n., 2012. towards an integrated network of coral immune mechanisms. proc. r. soc. b biol. sci. 279:4106–14. palmer, c. v., mydlarz, l.d., willis, b.l., 2008. evidence of an inflammatory-like response in non-normally pigmented tissues of two scleractinian corals. proc. r. soc. b biol. sci. 275:2687–93. palmer, c. v., traylor-knowles, n.g., 2018. cnidaria: anthozoans in the hot seat, in: cooper, e.l. (ed.), advances in comparative immunology. springer international publishing, cham, switzerland. p. 51–93. parisi, m.g., parrinello, d., stabili, l., cammarata, m., 2020. cnidarian immunity and the repertoire of defense mechanisms in anthozoans. biology (basel). 11:283. park, e., hwang, d.s., lee, j.s., song, j.i., seo, t.k., won, y.j., 2012. estimation of divergence times in cnidarian evolution based on mitochondrial protein-coding genes and the fossil record. mol. phylogenet. evol. 62:329–45. parker, h.j., krumlauf, r., 2017. segmental arithmetic: summing up the hox gene regulatory network for hindbrain development in chordates. wiley interdiscip. rev. dev. biol. 6:e286. patton, w.k., 1994. distribution and ecology of animals associated with branching corals (acropora spp.) from the great barrier reef, australia. bull. mar. sci. 55:193–211. patwa, a., thiéry, a., lombard, f., lilley, m.k.s., boisset, c., bramard, j.f., et al., 2015. accumulation of nanoparticles in “jellyfish” mucus: a bio-inspired route to decontamination of nano-waste. sci. rep. 5:11387. pauly, d., graham, w., libralato, s., morissette, l., deng palomares, m.l., 2009. jellyfish in ecosystems, online databases, and ecosystem models. hydrobiologia. 616:67–85. pearson, r., tellam, r., xu, b., zhao, z., willcox, m., kongsuwan, k., 2011. isolation, biochemical characterization and anti-adhesion property of mucin from the blue blubber jellyfish (catostylus mosaicus). available from: https://bioscipublisher.com/index.php/bm/article/view/165 perissinotto, r., pakhomov, e.a., 1998. contribution of salps to carbon flux of marginal ice zone of the lazarev sea, southern ocean. mar. biol. 131:25–32. peters, e.c., meyers, p.a., yevich, p.p., blake, n.j., 1981. bioaccumulation and histopathological effects of oil on a stony coral. mar. pollut. bull. 12:333–9. petralia, r.s., mattson, m.p., yao, p.j., 2014. aging and longevity in the simplest animals and the quest for immortality. ageing res. rev. 16:66–82. piggot, a.m., fouke, b.w., sivaguru, m., sanford, r.a., gaskins, h.r., 2009. change in zooxanthellae and mucocyte tissue density as an adaptive response to environmental stress by the coral, montastraea annularis. mar. biol. 156:2379–89. pitt, k.a., welsh, d.t., condon, r.h., 2009. influence of jellyfish blooms on carbon, nitrogen and phosphorus cycling and plankton production. hydrobiologia. 616:133–49. porporato, e.m.d., lo giudice, a., michaud, l., de domenico, e., spanò, n., 2013. diversity and antibacterial activity of the bacterial communities associated with two mediterranean sea pens, pennatula phosphorea and pteroeides spinosum (anthozoa: octocorallia). microb. ecol. 66:701–14. purcell, j.e., 1989. predation on fish larvae and eggs by the hydromedusa aequorea victoria at a herring spawning ground in british columbia. can. j. fish. aquat. sci. 46:1415–27. purcell, j.e., 2005. climate effects on formation of jellyfish and ctenophore blooms: a review. j. mar. biol. assoc. 85:461–76. purcell, j.e., 2012. jellyfish and ctenophore blooms coincide with human proliferations and environmental perturbations. ann. rev. mar. sci. 4:209–35. purcell, j.e., arai, m.n., 2001. interactions of pelagic cnidarians and ctenophores with fish: a review. hydrobiologia. 451:27–44. purcell, j.e., clarkin, e., doyle, t.k., 2012. foods of velella velella (cnidaria: hydrozoa) in algal rafts and its distribution in irish seas. hydrobiologia. 690:47–55. purcell, j.e., sturdevant, m. v., 2001. prey selection and dietary overlap among zooplanktivorous jellyfish and juvenile fishes in prince william sound, alaska. mar. ecol. prog. ser. 210:67–83. ramondenc, s., eveillard, d., guidi, l., lombard, f., delahaye, b., 2020. probabilistic modeling to estimate jellyfish ecophysiological properties and size distributions. sci. rep. 10:1–13. reshef, l., koren, o., loya, y., zilber-rosenberg, i., rosenberg, e., 2006. the coral probiotic hypothesis. environ. microbiol. 8:2068–73. reverter, m., tapissier-bontemps, n., lecchini, d., banaigs, b., sasal, p., 2018. biological and ecological roles of external fish mucus: a review. fishes. 3:41. richardson, a.j., bakun, a., hays, g.c., gibbons, m.j., 2009. the jellyfish joyride: causes, consequences and management responses to a more gelatinous future. trends ecol. evol. 24:312–22. riegl, b., branch, g.m., 1995. effects of sediment on the energy budgets of four scleractinian (bourne 1900) and five alcyonacean (lamouroux 1816) corals. j. exp. mar. bio. ecol. 186:259–75. rinkevich, b., wolodarsky, z., loya, y., 1991. coral-crab association: a compact domain of a multilevel trophic system. hydrobiologia. 216:279–84. ritchie, k.b., 2006. regulation of microbial populations by coral surface mucus and mucus-associated bacteria. mar. ecol. prog. ser. 322:1–14. ritchie, k.b., smith, g.w., 2004. microbial communities of coral surface mucopolysaccharide layers. coral health and disease. p. 259–64. rivera-ortega, j., thomé, p.e., 2018. contrasting antibacterial capabilities of the surface mucus layer from three symbiotic cnidarians. front. mar. sci. 5:392. no nco mm er cia l u se on ly s. savoca et al.92 rodrigues, l.j., grottoli, a.g., 2007. energy reserves and metabolism as indicators of coral recovery from bleaching. limnol. oceanogr. 52:1874–82. rogers, c.s., 1990. responses of coral reefs and reef organisms to sedimentation. mar. ecol. prog. ser. 62:185–202. rohwer, f., breitbart, m., jara, j., azam, f., knowlton, n., 2001. diversity of bacteria associated with the caribbean coral montastraea franksi. coral reefs. 20:85–91. romano, s.l., palumbi, s.r., 1996. evolution of scleractinian corals inferred from molecular systematics. science. 271:640–2. rosenberg, e., koren, o., reshef, l., efrony, r., zilber-rosenberg, i., 2007. the role of microorganisms in coral health, disease and evolution. nat. rev. microbiol. 5:355–62. rosenberg, e., sharon, g., zilber-rosenberg, i., 2009. the hologenome theory of evolution contains lamarckian aspects within a darwinian framework. environ. microbiol. 11:2959–62. rosenberg, e., zilber-rosenberg, i., 2011. symbiosis and development: the hologenome concept. birth defects res. part c embryo today rev. 93:56–66. rossi, s., ribes, m., coma, r., gili, j.m., 2004. temporal variability in zooplankton prey capture rate of the passive suspension feeder leptogorgia sarmentosa (cnidaria: octocorallia), a case study. mar. biol. 144:89–99. rottini sandrini, l., avian, m., 1991. reproduction of pelagia noctiluca in the central and northern adriatic sea. hydrobiologia. 216:197–202. rublee, p.a., lasker, h.r., gottfried, m., roman, m.r., 1980. production and bacterial colonization of mucus from the soft coral briarium asbestinum. bull. mar. sci. 30:888–93. ruzicka, j., brodeur, r.d., cieciel, k., decker, m.b., 2020. examining the ecological role of jellyfish in the eastern bering sea. ices j. mar. sci. 77:791–802. rypien, k.l., ward, j.r., azam, f., 2010. antagonistic interactions among coral-associated bacteria. environ. microbiol. 12:28–39. sadd, b.m., schmid-hempel, p., 2009. principles of ecological immunology. evol. appl. 2:113–21. santos, g.s., amaral, f.d., sassi, c.f.c., schwamborn, r., 2016. response of the zooxanthellae of palythoa caribaeorum (cnidaria: zoanthidea) to different environmental conditions in coastal and oceanic ecosystems of the tropical atlantic. helgol. mar. res. 70:2. santos, m.e.a., kitahara, m.v., lindner, a., reimer, j.d., 2016. overview of the order zoantharia (cnidaria: anthozoa) in brazil. mar. biodivers. 46:547–59. sanz-martín, m., pitt, k.a., condon, r.h., lucas, c.h., novaes de santana, c., duarte, c.m., 2016. flawed citation practices facilitate the unsubstantiated perception of a global trend toward increased jellyfish blooms. glob. ecol. biogeogr. 25:1039–49. savoca, s., lo giudice, a., papale, m., mangano, s., caruso, c., spanò, n., et al., 2019. antarctic sponges from the terra nova bay (ross sea) host a diversified bacterial community. sci. rep. 9:16135. schaub, j., hunt, b.p. v, pakhomov, e.a., holmes, k., lu, y., quayle, l., 2018. using unmanned aerial vehicles (uavs) to measure jellyfish aggregations. mar. ecol. prog. ser. 591:29–36. schöttner, s., hoffmann, f., wild, c., rapp, h.t., boetius, a., ramette, a., 2009. inter-and intra-habitat bacterial diversity associated with cold-water corals. isme j. 3:756–9. schuhmacher, h., 1977. ability in fungiid corals to overcome sedimentation. 3rd international coral reef symposium. available from: http://koha.ideam.gov.co/cgi-bin/koha/opac-detail. pl?biblionumber=4434457 sebens, k.p., 1982. the limits to indeterminate growth: an optimal size model applied to passive suspension feeders. ecology. 63:209–22. shanks, a., graham, w., 1988. chemical defense in a scyphomedusa. mar. ecol. prog. ser. 45:81–6. sharp, k.h., distel, d., paul, v.j., 2012. diversity and dynamics of bacterial communities in early life stages of the caribbean coral porites astreoides. isme j. 6:790–801. sheldon, b.c., verhulst, s., 1996. ecological immunology: costly parasite defences and trade-offs in evolutionary ecology. trends ecol. evol. 11:317–21. sheridan, c., grosjean, p., leblud, j., palmer, c. v., kushmaro, a., eeckhaut, i., 2014. sedimentation rapidly induces an immune response and depletes energy stores in a hard coral. coral reefs. 33:1067–76. shick, j.m., 1991. energy metabolism and respiratory gas exchange, in: a functional biology of sea anemones. springer, amsterdam, the netherlands. p. 119–173. shinzato, c., shoguchi, e., kawashima, t., hamada, m., hisata, k., tanaka, m., et al., 2011. using the acropora digitifera genome to understand coral responses to environmental change. nature. 476:320–3. shnit-orland, m., kushmaro, a., 2009. coral mucus-associated bacteria: a possible first line of defense. fems microbiol. ecol. 67:371–80. siddall, m.e., martin, d.s., bridge, d., desser, s.s., cone, d.k., 1995. the demise of a phylum of protists: phylogeny of myxozoa and other parasitic cnidaria. j. parasitol. 81:961–7. siva-jothy, m.t., thompson, j.j.w., 2002. short-term nutrient deprivation affects immune function. physiol. entomol. 27:206–12. sleigh, m.a., 1989. adaptations of ciliary systems for the propulsion of water and mucus. comp. biochem. physiol. a. comp. physiol. 94:359–64. smith, v.j., desbois, a.p., dyrynda, e.a., 2010. conventional and unconventional antimicrobials from fish, marine invertebrates and micro-algae. mar. drugs. 8:1213-62. sokolow, s., 2009. effects of a changing climate on the dynamics of coral infectious disease: a review of the evidence. dis. aquat. organ. 87:5–18. sommer, u., lengfellner, k., 2008. climate change and the timing, magnitude, and composition of the phytoplankton spring bloom. glob. chang. biol. 14:1199–208. soetaert, k., mohn, c., rengstorf, a., grehan, a., van oevelen, d., 2016. ecosystem engineering creates a direct nutritional link between 600-m deep cold-water coral mounds and surface productivity. sci. rep. 6:35057. stabili, l., rizzo, l., basso, l., marzano, m., fosso, b., pesole, g., piraino, s., 2020. the microbial community associated with rhizostoma pulmo: ecological significance and potential consequences for marine organisms and human health. mar. drugs. 18:437. stabili, l., schirosi, r., parisi, m.g., piraino, s., cammarata, m., no nco mm er cia l u se on ly mucus secretions in cnidarian, an ecological, adaptive and evolutive tool 93 2015. the mucus of actinia equina (anthozoa, cnidaria): an unexplored resource for potential applicative purposes. mar. drugs. 13:5276–96. stafford-smith, m.g., 1993. sediment-rejection efficiency of 22 species of australian scleractinian corals. mar. biol. 115:229–43. stafford-smith, m.g., ormond, r.f.g., 1992. sediment-rejection mechanisms of 42 species of australian scleractinian corals. mar. freshw. res. 43:683–705. steinberg, d.k., landry, m.r., 2017. zooplankton and the ocean carbon cycle. ann. rev. mar. sci. 9:413–44. suchman, c.l., daly, e.a., keister, j.e., peterson, w.t., brodeur, r.d., 2008. feeding patterns and predation potential of scyphomedusae in a highly productive upwelling region. mar. ecol. prog. ser. 358:161–72. tanaka, y., miyajima, t., umezawa, y., hayashibara, t., ogawa, h., koike, i., 2009. net release of dissolved organic matter by the scleractinian coral acropora pulchra. j. exp. mar. bio. ecol. 377:101–6. tarrant, a.m., 2007. hormonal signaling in cnidarians: do we understand the pathways well enough to know whether they are being disrupted? ecotoxicology. 16:5–13. technau, u., steele, r.e., 2011. evolutionary crossroads in developmental biology: cnidaria. development. 138: 1447–58. thompson, j.h., shinn, e.a., bright, t.j., 1980. chapter 16 effects of drilling mud on seven species of reef-building corals as measured in the field and laboratory. elsevier oceanogr. ser. 27:433–53. tinta, t., klun, k., herndl, g.j., 2021. the importance of jellyfish–microbe interactions for biogeochemical cycles in the ocean. limnol. oceanogr. 66:2011–32. tinta, t., kogovšek, t., klun, k., malej, a., herndl, g.j., turk, v., 2019. jellyfish-associated microbiome in the marine environment: exploring its biotechnological potential. mar. drugs. 17:94. tinta, t., kogovšek, t., malej, a., turk, v., 2012. jellyfish modulate bacterial dynamic and community structure. plos one. 7:e39274. tinta, t., kogovšek, t., turk, v., shiganova, t.a., mikaelyan, a.s., malej, a., 2016. microbial transformation of jellyfish organic matter affects the nitrogen cycle in the marine water column—a black sea case study. j. exp. mar. bio. ecol. 475:19–30. tinta, t., malej, a., kos, m., turk, v., 2010. degradation of the adriatic medusa aurelia sp. by ambient bacteria. hydrobiologia. 645:179-91. titelman, j., riemann, l., sørnes, t.a., nilsen, t., griekspoor, p., båmstedt, u., 2006. turnover of dead jellyfish: stimulation and retardation of microbial activity. mar. ecol. prog. ser. 325:43–58. tremblay, p., naumann, m.s., sikorski, s., grover, r., ferrierpagès, c., 2012. experimental assessment of organic carbon fluxes in the scleractinian coral stylophora pistillata during a thermal and photo stress event. mar. ecol. prog. ser. 453:63–77. turk, v., fortič, a., kramar, m.k., žnidarič, m.t., štrus, j., kostanjšek, r., malej, a., 2021. observations on the surface structure of aurelia solida (scyphozoa) polyps and medusae. diversity. 13:244. vacelet, e., thomassin, b.a., 1991. microbial utilization of coral mucus in long term in situ incubation over a coral reef. hydrobiologia. 211:19–32. van oppen, m.j.h., gates, r.d., blackall, l.l., cantin, n., chakravarti, l.j., chan, w.y., et al., 2017. shifting paradigms in restoration of the world’s coral reefs. glob. chang. biol. 23:3437–48. wagner, d., luck, d.g., toonen, r.j., 2012. the biology and ecology of black corals (cnidaria: anthozoa: hexacorallia: antipatharia). adv. mar. biol. 63:67-132. wasserthal, l.t., wasserthal, w., 1973. ökologische bedeutung der schleimsekretion bei den planula-larven der hydroidengattung eudendrium. mar. biol. 22:341–345. weber, m., de beer, d., lott, c., polerecky, l., kohls, k., abed, r.m.m., et al., 2012. mechanisms of damage to corals exposed to sedimentation. proc. natl. acad. sci. u. s. a. 109:e1558–67. webster, n.s., bourne, d., 2007. bacterial community structure associated with the antarctic soft coral, alcyonium antarcticum. fems microbiol. ecol. 59:81–94. weiland-bräuer, n., neulinger, s.c., pinnow, n., künzel, s., baines, j.f., schmitz, r.a., 2015. composition of bacterial communities associated with aurelia aurita changes with compartment, life stage, and population. appl. environ. microbiol. 81:6038–52. weiland-bräuer, n., pinnow, n., langfeldt, d., roik, a., güllert, s., chibani, c.m., et al., 2020. the native microbiome is crucial for offspring generation and fitness of aurelia aurita. mbio. 11:e02336-20. welsh, d.t., dunn, r.j.k., meziane, t., 2009. oxygen and nutrient dynamics of the upside down jellyfish (cassiopea sp.) and its influence on benthic nutrient exchanges and primary production. hydrobiologia. 635:351–62. west, e.j., pitt, k.a., welsh, d.t., koop, k., rissik, d., 2009. top-down and bottom-up influences of jellyfish on primary productivity and planktonic assemblages. limnol. oceanogr. 54:2058–71. wild, c., huettel, m., klueter, a., kremb, s.g., rasheed, m.y.m., jørgensen, b.b., 2004. coral mucus functions as an energy carrier and particle trap in the reef ecosystem. nature. 428:66–70. wild, c., laforsch, c., huettel, m., 2006. detection and enumeration of microbial cells within highly porous calcareous reef sands. mar. freshw. res. 57:415–20. wild, c., naumann, m., niggl, w., haas, a., 2010. carbohydrate composition of mucus released by scleractinian warm-and cold-water reef corals. aquat. biol. 10:41–5. wild, c., woyt, h., huettel, m., 2005. influence of coral mucus on nutrient fluxes in carbonate sands. mar. ecol. prog. ser. 287:87–98. wolenski, f.s., chandani, s., stefanik, d.j., jiang, n., chu, e., finnerty, j.r., gilmore, t.d., 2011. two polymorphic residues account for the differences in dna binding and transcriptional activation by nf-jb proteins encoded by naturally occurring alleles in nematostella vectensis. j. mol. evol. 73:325–36. wooldridge, s.a., 2009. a new conceptual model for the enhanced release of mucus in symbiotic reef corals during “bleaching” conditions. mar. ecol. prog. ser. 396:145–52. wooldridge, s.a., 2017. instability and breakdown of the coral– no nco mm er cia l u se on ly s. savoca et al.94 algae symbiosis upon exceedence of the interglacial pco2 threshold (>260 ppmv): the “missing” earth-system feedback mechanism. coral reefs. 36:1025–37. work, t., meteyer, c., 2014. to understand coral disease, look at coral cells. ecohealth. 11:610–18. yonge, c.m., 1930. studies on the physiology of corals iv. the structure distribution and physiology of the zooxanthellae. available from: https://archive.org/details/biostor-175049 zamzow, j.p., 2007. ultraviolet-absorbing compounds in the mucus of shallow-dwelling tropical reef fishes correlate with environmental water clarity. mar. ecol. prog. ser. 343:263–271. zapata, f., goetz, f.e., smith, s.a., howison, m., siebert, s., church, s.h., et al., 2015. phylogenomic analyses support traditional relationships within cnidaria. plos one. 10:e0139068. zhao, y., parry, l.a., vinther, j., dunn, f.s., li, y., wei, f., 2021. an early cambrian polyp reveals an anemone-like ancestor for medusozoan cnidarians. available from: https://www. biorxiv.org/content/10.1101/2021.12.24.474121v1 no nco mm er cia l u se on ly layout 1 introduction toxigenic cyanobacteria are one of the major health risks associated with water resources, in europe and beyond. cyanobacteria can produce a wide range of potent toxins (cyanotoxins) with adverse health effects on humans and animals exposed e.g., via drinking water, aquaculture and recreation. the hepatotoxic peptide toxins (microcystins: mcs) occur annually in most countries and this cyanotoxin family is considered a major concern in the european context. mcs have been identified as undisputed health hazards by the world health organization (who). emerging cyanotoxins are also of concern in europe; e.g., cylindrospermopsin (cyn) which was viewed as being limited to (sub)tropical waters but has been recently found in several european countries. it has been postulated that such emerging cyanotoxins could be related to invasive cyanobacterial species that are colonizing european waters. recent research has also demonstrated a definite relationship, on a global scale, between temperature and the dominance of cyanobacteria in water bodies (o’neil et al., 2012). advanced methods have already been developed in european laboratories to detect, identify and quantitate cyanotoxins. although iso 20179:2005 introduced a standard method to determine mcs, other methods for cyanotoxin analysis are not harmonized across europe. techniques including liquid chromatography mass specadvances in oceanography and limnology, 2017; 8(1): 161-178 article doi: 10.4081/aiol.2017.6429 this work is licensed under a creative commons attribution-noncommercial 4.0 international license (cc by-nc 4.0). toxic cyanobacteria and cyanotoxins in european waters recent progress achieved through the cyanocost action and challenges for further research jussi meriluoto,1,2* ludek blaha,3 gorenka bojadzija,4 myriam bormans,4 luc brient,4 geoffrey a. codd,5,6 damjana drobac,2 elisabeth j. faassen,7 jutta fastner,8 anastasia hiskia,9 bastiaan w. ibelings,10 triantafyllos kaloudis,11 mikolaj kokocinski,12 rainer kurmayer,13 dijana pantelić,2 antonio quesada,14 nico salmaso,15 nada tokodi,2 theodoros m. triantis,9 petra m. visser,16 zorica svirčev1,2 1biochemistry, faculty of science and engineering, åbo akademi university, tykistökatu 6a, 20520 turku, finland; 2laboratory for paleoenvironmental reconstruction, faculty of sciences, university of novi sad, trg dositeja obradovica 2, 21000 novi sad, serbia; 3recetox, faculty of science, masaryk university, kamenice 5, cz62500 brno, czech republic; 4umr ecobio 6553 cnrs université de rennes, 1 campus de beaulieu, rennes, france; 5biological and environmental sciences, university of stirling, stirling fk9 4la, scotland; 6college of life sciences, university of dundee, dow street, dundee dd1 5eh, scotland; 7aquatic ecology & water quality management group, wageningen university, p.o. box 47, wageningen 6700 dd, the netherlands; 8section drinking water resources and water treatment, german environment agency, corrensplatz 1, 14195 berlin, germany; 9institute of nanoscience & nanotechnology, national center for scientific research “demokritos”, neapoleos 10, ag. paraskevi 153 10, athens, greece; 10department f.-a. forel for environmental and aquatic sciences and institute for environmental sciences, university of geneva, 66 boulevard carl-vogt, 1211 geneva 4, switzerland; 11department of water quality control, athens water supply and sewerage company, laodikias 29, athens, greece; 12department of hydrobiology, faculty of biology, adam mickiewicz university, wieniawskiego 1, 61-712 poznań, poland; 13research institute for limnology, university of innsbruck, mondseestrasse 9, 5310 mondsee, austria; 14department of biology, universidad autónoma de madrid, 28049 madrid, spain; 15department of sustainable agrosystems and bioresources, research and innovation centre, fondazione edmund mach, via e. mach 1, 38010 san michele all’adige, tn, italy; 16department of aquatic microbiology, institute for biodiversity and ecosystem dynamics, university of amsterdam, p.o. box 94248, 1090 ge amsterdam, the netherlands *corresponding author: jussi.meriluoto@abo.fi abstract this review aims to summarise the outcomes of some recent european research concerning toxic cyanobacteria and cyanotoxins, with an emphasis on developments within the framework of the cyanocost action: cost action es1105, cyanobacterial blooms and toxins in water resources: occurrence, impacts and management. highlights of the action include phycological and ecological studies, development of advanced techniques for cyanotoxin analysis, elucidation of cyanotoxin modes of action, management techniques to reduce cyanobacterial mass development, and research on methods and practices for cyanotoxin removal during drinking water treatment. the authors have identified a number of gaps in knowledge. proposed directions for future research on toxic cyanobacteria and cyanotoxins are also discussed. key words: cyanobacteria; cyanotoxins; europe; trends; gaps of knowledge. received: 29 november 2016. accepted: 7 march 2017.n on -co mm er cia l u se on ly j. meriluoto et al.162 trometry (lc-ms) have enabled cyanotoxin detection and identification at very low concentrations. rapid immunoassay methods, for on-site use, are also being developed in europe, and merit dissemination with verification, as kits, to increase capabilities in regions without hightechnology resources. multiple techniques (proactive and reactive) have already been used to control cyanobacterial blooms in europe, with varying degrees of success. control measures for cyanobacterial and cyanotoxin management have been developed, and these range from catchment and basin-scale management, to drinking water treatment methods for the removal of cyanobacterial cells and cyanotoxins, including sorption, filtration and advanced oxidation processes. an important part of almost any research is end-user involvement. this is especially true in research concerning public health questions including the research activities related to cyanocost. occurrence of cyanobacteria and cyanotoxins, and methods for monitoring and analysis the continuing eutrophication of many aquatic ecosystems and water resources in europe is causing concern because the deterioration of water quality and potentially toxic blooms of cyanobacteria threaten human and animal health. throughout the cyanocost action it was shown that toxic cyanobacteria and various cyanotoxins are commonly found in all parts of europe. however, it is difficult to make firm conclusions about any trends in the occurrence of toxic cyanobacteria and cyanotoxins as monitoring has only recently been introduced in some parts of europe. there has been very little systematic cyanotoxin monitoring of waters in most countries as the cyanotoxin analyses have been primarily carried out by academic partners involved in specific research projects or, in some cases, by public or private water utilities abstracting water from potentially high-risk sources. it is also likely that there has been a certain bias towards the cyanotoxins for which the analytical methods are more widely established, i.e. for the mcs, and the full spectrum of cyanotoxins present in europe still remains to be comprehensively assessed. what can be concluded without doubt is that those laboratories that have been involved in cyanotoxin analyses for a longer time (some since the 1980s) continue to find mcs in many water environments monitored since that time. in addition, further cyanobacterial hazards (invasive species and emerging/less-studied toxins) are being discovered. the joint activity of cyanocost and the netlake cost action (es1201), the european multilake survey (emls) conducted in 2015 will be able to provide a snapshot of the cyanobacterial/cyanotoxin situation in a wide range of european waters when the analytical work concentrated in selected laboratories has been finished. it would be highly desirable to be able to continue and expand the monitoring activities initiated within the emls and thus get a clearer picture of any trends in cyanotoxin occurrence and abundance in europe. hepatotoxic mcs appear to be the most common and to have the greatest known impacts on human health and activities among the known cyanotoxins in aquatic ecosystems including lakes and rivers across europe. major producers of mcs include cyanobacteria from the genus microcystis and members of other cyanobacterial genera including planktothrix, nostoc, anabaena/dolichospermum and anabaenopsis (merel et al., 2013a, 2013b; bernard et al., 2017). nodularin (nod) hepatotoxins are mostly known from the baltic sea. however, recently nodularin was found also in aquatic ecosystems in turkey (mazur-marzec et al., 2013; sahindokuyucu kocasari et al., 2015). moreover, nodularia spumigena mertens ex bornet & flahault, a nod-producer, was recently observed not only in brackish waters but also in freshwater lakes (akcaalan et al., 2009). acutely-acting neurotoxins are much less common than hepatotoxins but are still found in a large variety of water ecosystems. anatoxin-a (antx-a) and/or homoanatoxin-a (hantx-a) have been detected in lakes and rivers in the uk, ireland, france, netherlands, germany, italy, denmark, portugal, and poland (edwards et al., 1992; furey et al., 2003; gugger et al., 2005; messineo et al., 2009; osswald et al., 2009; ballot et al., 2010; faassen et al., 2012b; kobos et al., 2013). these toxins are associated with several cyanobacteria including, among others, species of anabaena, dolichospermum and aphanizomenon (for an updated list see bernard et al., 2017). a less toxic analogue, dihydroanatoxin-a was isolated from different strains of oscillatoria sp., phormidium sp. and cylindrospermum stagnale bornet & flahault (méjean et al., 2014). evidence for the potent acetylcholinesterase inhibitor anatoxin-a(s) (onodera et al., 1997) has been found more rarely in europe: to date, in denmark (henriksen et al., 1997) and scotland (devic et al., 2002). saxitoxins (stxs) have also been rarely reported from european freshwaters. these potent neurotoxins were observed in several countries including france, finland, and spain associated mostly with the occurrence of aphanizomenon spp. (rapala et al., 2005; ledreux et al., 2010; cirés et al., 2014). the cytotoxin and genotoxin cyn was until recently only known from warmer regions. during the last decade, it has been detected in several european countries. a further discussion about cyn is found later in this section. in addition to these three major groups of cyanotoxins, i.e., hepatotoxins, neurotoxins and cytotoxins, there is abundant toxicological evidence for further unidentified no nco mm er cia l u se on ly toxic cyanobacteria in europe research progress through cyanocost 163 cyanobacterial bioactive/toxic compounds in aquatic ecosystems in europe (elersek et al., 2017). besides new toxins and indications of a widening occurrence of known toxins, future research may reveal new environments affected with toxic/noxious cyanobacteria and their toxins. additional threats to aquatic ecosystems may be caused by alien and invasive species that are expanding their geographical range (kokociński et al., 2017). cylindrospermopsis raciborskii (woloszynska) seenayya & subba raju is an example of a cyanobacterium that has increased its distribution from tropical and subtropical regions towards the temperate zone where it is now widespread. this species is known for the production of cyn and stxs in warmer regions but, so far, this was not confirmed in europe. recent studies, however, revealed the ability of this cyanobacterium to produce unknown metabolites causing toxic effects (poniedziałek et al., 2015). among other alien species that occur more and more frequently in europe are chrysosporum bergii (ostenfeld) e. zapomělová, o. skácelová, p. pumann, r. kopp & e. janeček, chrysosporum ovalisporum (forti) e. zapomělová, o. skácelová, p. pumann, r. kopp & e. janeček, and sphaerospermopsis aphanizomenoides (forti) zapomělová, jezberová, hrouzek, hisem, reháková & komárková. these are known cyn-producers (with the exception of the last one the toxicity of which remains debatable) (quesada et al., 2006; kaštovskỳ et al., 2010; koreivienė and kasperovičenė, 2011) and an increasing distribution and abundance of these species may contribute to additional toxic stress in european freshwaters. already before 2012, i.e., the start of the cyanocost action, mcs and antxs had been involved in poisoning episodes affecting animal and/or human health in europe but no health effects caused by cyn had been observed/reported. the (apparent or true) increasing occurrence of cyn in europe is likely to change this picture, and health effects caused by cyn can be expected especially in southern and central europe, the source of most of the reports concerning cyn. in addition, there have been cases of reported poisonings in european waters where unknown/new cyanotoxins have been inferred but a more exact identification was lacking. health effects and poisoning episodes related to cyanotoxin-contaminated drinking and recreational waters were documented during the lifetime of cyanocost. however, the number of reported cases was relatively low. this may indicate an increased awareness and implementation of better monitoring and management systems for toxic cyanobacteria. there have also been improvements in analytical capability for cyanotoxin detection in many countries, partly thanks to increased international collaboration and the spreading of know-how. a possible reason for fewer reports in the scientific literature may be the fact that it is increasingly difficult to get cyanotoxin episodes published unless something really dramatic or novel is involved. however, as shown by the užice case reported here (see fate, impact and health effects), recognition of the occurrence and impacts of toxigenic cyanobacteria remains inadequate in europe, and cyanotoxins continue to cause serious health effects, not only in individuals but also at the population level. the lack of international and national regulatory guidelines for most cyanotoxins in drinking water is a serious problem which needs to be addressed urgently and which was highlighted also in the užice case. while modern sensitive analytical equipment can relatively easily comply with the most demanding monitoring needs, the lack of regulatory guidelines for most cyanotoxins is likely to give the cyanotoxins a lower priority in monitoring. also, the lack of commercial analytical reference materials is problematic. among the cyanotoxins, there are two toxins of growing importance in european waters and elsewhere: cyn and β-n-methylamino-l-alanine (bmaa). cyn is a potent cytotoxic and genotoxic cyanotoxin initially detected in australia (ohtani et al., 1992; saker and neilan, 2001). later, it was reported in new zealand (stirling and quilliam, 2001), thailand (li et al., 2001) and other tropical and subtropical regions (rzymski and poniedziałek, 2014). cyn was named after the first identified producer c. raciborskii being involved in human poisoning and an outbreak of hepatoenteritis among the aboriginal community in australia. nowadays cyn has been detected more widely and there are more than 12 known producers (de la cruz et al., 2013). in europe, this toxin was first reported in germany in 2000 in two shallow lakes in the eastern part of brandenburg (fastner et al., 2003). further, in hungary in 2002 (kiss et al., 2002), italy and spain in 2004 (manti et al., 2005; quesada et al., 2006), finland, france, the czech republic and poland in 2006 (spoof et al., 2006; bláhová et al., 2009; brient et al., 2009; kokociński et al., 2009). in europe, cyanobacteria capable of producing cyn have most often been found in the genera aphanizomenon, anabaena/dolichospermum, chrysosporum and oscillatoria (kokociński et al., 2013; rzymski and poniedziałek, 2014; bernard et al., 2017). more potential producers of cyn have been revealed lately including dolichospermum mendotae (w. trelease) p. wacklin, l. hoffmann & j. komárek in turkey and the terrestrial cyanobacterium hormoscilla pringsheimii anagnostidis & komárek (akcaalan et al., 2014; bohunická et al., 2015). moreover, gkelis and zaoutsos (2014) for the first time reported cyn and a c. raciborskii bloom producing stx in greece but a definitive identification of the cyn producer could not be made. in europe, in contrast to warmer regions, thus only non-cyn producing populations of c. raciborskii have been so far identified except for a recent report from serbia (đorđević et al., 2015). the latter observation however is disputable no nco mm er cia l u se on ly j. meriluoto et al.164 as no studies on isolated cultures of this cyanobacterium were performed. complementary studies have been recently conducted on the toxicity of cyn (guzmán-guillén et al., 2015; sierosławska et al., 2015;), mechanisms of toxicity (poniedziałek et al., 2014), biodegradation (dziga et al., 2016) and biological role and allelopathy (b-béres et al., 2015). these works have contributed valuable knowledge on these aspects but at the same time have revealed many gaps in understanding of the role(s) of cyn in aquatic ecosystems and its impact on their functioning (sukenik et al., 2015). future studies are needed also on factors regulating biosynthesis of this toxin, its biodegradation by bacterial communities and other potential producers including benthic and terrestrial cyanobacteria. 7-deoxy-cyn and 7-epi-cyn are naturally occurring congeners of cyn (reviewed by de la cruz et al., 2013). different aphanizomenon sp. strains obtained from german lakes were confirmed as producers of 7-deoxy-cyn while strains of chr. ovalisporum and c. raciborskii are known to produce also 7-epi-cyn (rzymski and poniedziałek, 2014; cirés and ballot, 2016). recently two new analogs, 7-deoxy-desulfo-cyn and 7-deoxy-desulfo-12-acetylcyn, were isolated from a thai strain of c. raciborskii (wimmer et al., 2014). in the period of 2012-2016, important steps towards a better understanding of the presence of the neurotoxin bmaa in surface waters have been made. the presence of bmaa in a diversity of cyanobacterial strains was first described in 2005 (cox et al., 2005), and subsequent studies reported conflicting results (reviewed by faassen, 2014). differences in analytical approaches were assumed to be one of the main causes of these diverging results, and in 2012, two independent studies recommended suitable analytical methods for bmaa determination (cohen 2012; faassen et al., 2012a). although some analytical challenges remain (lage et al., 2015; faassen et al., 2016), a more consistent picture of the presence of bmaa in aquatic systems appears when only the studies that used recommended analytical approaches (cohen, 2012; duncan, 2012; faassen et al., 2012a) are considered (faassen, 2014). in europe, bmaa has frequently been detected in aquatic organisms sometimes intended for human consumption, including fish, molluscs and other seafood, at or below the low µg g–1 dw level (faassen, 2014; jiang et al., 2014b; réveillon et al., 2016b). similar levels are found in some phytoplankton samples or strains (faassen 2014). while at first cyanobacteria were hypothesised to be the source of bmaa in aquatic systems, bmaa has now also been detected in diatoms (jiang et al., 2014a; lage et al., 2015; réveillon et al., 2015), and a diatom strain has been shown to contain bmaa during different growth phases (réveillon et al., 2016a). recently, a major fraction of bmaa was found to be present in a soluble, bound form in a diversity of samples (faassen et al., 2016; rosén et al., 2016), where traditionally bmaa was considered to be present either as a free molecule or in a protein-associated form (murch et al., 2004). only if more consensus is created on the presence of bmaa in aquatic systems, and human exposure levels can be assessed with confidence, will it be possible to determine whether bmaa does indeed play a role in the globally-occurring neurodegenerative diseases alzheimer’s, parkinsonism and amyotrophic lateral sclerosis, as is currently hypothesized (bradley and mash, 2009). the traditional identification of cyanobacteria has been based on light microscopy and this technique continues to dominate in most laboratories throughout europe as documented in the recent handbook of cyanobacterial monitoring and cyanotoxin analysis (meriluoto et al., 2017). however, the morphological species concept is changing in the microbial world, and molecular tools are impacting on this regarding cyanobacteria. the rapidly growing body of information from molecular biological studies (see the text about the second handbook later in this section, kurmayer et al., 2017) is also very useful in distinguishing between potentially toxigenic and non-toxigenic cyanobacterial strains but which may have identical morphological features. advances are being made in cyanotoxin screening and trace analysis. commercial enzyme-linked immunoassays (elisas) are being applied for the screening for some cyanotoxins, as are protein phosphatase inhibition assays for the detection of mcs and nods. however, the need for users to themselves verify the specificity of these methods, taking into account false positives and responses to cyanotoxin detoxification products, is emphasised. the most rapid and dramatic (in terms of sensitivity) advances have been made in instrumental analyses based on liquid chromatography-tandem mass spectrometry (lcms/ms). lc-ms/ms is extremely well suited for regulatory settings as it can target individual analytes and the methods can be fully validated. in addition, lc-ms can be used simultaneously for several classes of compounds. zervou et al. (2017) provide a good example of this multitoxin approach where 12mcs, nod, cyn, antx-a, and the marine algal toxins domoic acid and okadaic acid were concentrated by a dual-cartridge solid-phase extraction procedure and determined in a single lc-ms/ms run. one obvious complication with lc-ms/ms analyses is the need for reference materials which are currently not commercially available for all cyanotoxins. what has become increasingly evident during the last few years is that various bioassays are still needed in identifying new toxic agents produced by cyanobacteria. new bioassays based on cell lines or invertebrates have almost totally replaced the traditional mouse bioassay which was relied upon in the 1980s when the knowledge about cyanotoxins and their effects on mammalian/human health was no nco mm er cia l u se on ly toxic cyanobacteria in europe research progress through cyanocost 165 still developing. such bioassays can possibly detect novel bioactive, including potentially harmful, compounds. however, the broader application of bioassays in the monitoring of toxic cyanobacteria is still an open issue. the ongoing discussion among scientists and regulators aims to clarify important points such as causality between bioassay responses and toxin concentrations, bioassay specificity, action or regulatory limits, etc. in the past two decades, molecular biology has provided a major contribution to understanding cyanotoxin biosynthesis pathways (kurmayer et al., 2017). the elucidation of the pathways of mc and nod biosynthesis (tillett et al., 2000; moffitt and neilan, 2004) was followed by discovery of the pathways of cyn biosynthesis (shalev-alon et al., 2002; mihali et al., 2008), stx biosynthesis (kellmann et al., 2008) and antx-a biosynthesis (méjean et al., 2009). in general, cyanotoxin molecules are synthesized by non-ribosomal peptide synthesis (nrps) partly involving polyketide synthases (pks) and tailoring enzymes following a thiotemplate mechanism which is widely distributed among bacteria and fungi (fischbach and walsh, 2006). synthesis pathways have been explored by characterizing specific enzymatic steps in vitro (hicks et al., 2006), by experimental inactivation of core and tailoring enzymatic steps in vivo (christiansen et al., 2003), and by crystallization of heterologously-expressed enzymes (mazmouz et al., 2011). in parallel with gene cluster elucidation: i) various gene loci have been tested for their reliability to indicate the presence of a specific cyanotoxin biosynthesis pathway (jungblut and neilan, 2006); and ii) phylogenetic analysis of various gene loci has been used to investigate the inheritance of cyanotoxin biosynthesis pathways (rantala et al., 2004). this research has revealed specific gene loci indicative of mc/nod biosynthesis (rantala et al., 2004), antx-a biosynthesis (rantala-ylinen et al., 2011), and cyn biosynthesis (mihali et al., 2008), and preliminary candidate gene loci indicative of stx biosynthesis (kellmann et al., 2008). today it is understood that mc and stx biosynthesis pathways are both ancient and evolved about 2 billion years ago corresponding to the evolution of the heterocytous cyanobacteria as an adaptation to the oxygen-containing atmosphere (rantala et al., 2004; murray et al., 2011). consequently, due to an ancestral common origin for these cyanotoxins (mc/nod, stx), their present-day production is widely distributed among extant cyanobacteria and is likely to be more so than is currently recognised. this type of phylogenetic analysis has further suggested that, in comparison with horizontal gene transfer events, the vertical inheritance and ongoing cyanotoxin gene cluster loss processes are probably important in explaining the rather patchy distribution among cyanobacterial taxa. correspondingly, for cyn, phylogenetic congruence with cyanobacterial evolution and cyn biosynthesis has been reported (jiang et al., 2012). for antx-a biosynthesis, recent phylogenetic analysis has revealed significant homology within the antx-a biosynthesis pathway across cyanobacterial taxa (brown et al., 2016). current understanding of the molecular biology of cyanotoxin biosynthesis has been recently compiled into a molecular tools handbook which contains protocols describing methods to detect and to quantify toxigenic cyanobacteria in aquatic and terrestrial environments and in food supplements (kurmayer et al., 2017). this compilation contains protocols on sampling and cyanobacterial isolation and identification, and standard protocols on dna (mrna) extraction and pcr amplification of gene loci indicative of cyanotoxin biosynthesis. in addition to toxigenic cyanobacterial detection and quantification protocols, methods to analyze the regulation of cyanotoxin synthesis are also included. for quantitative analysis, the techniques used in qpcr have been collected. the compilation also includes traditional microarray and denaturing gradient gel electrophoresis protocols to characterize genetic community/population structure based on cyanotoxin synthesis gene fragments. high-throughput sequencing protocols have been included to aid the investigation of cyanobacterial community composition. finally, the application of molecular tools in the scientific literature has been collected and reviewed (kurmayer et al., 2017). in principal, molecular tools only can reveal the potential for cyanotoxin biosynthesis and cannot provide information about actual toxin concentrations or contents. in general, the scientific experience of the predictability of cyanotoxin occurrence qualitatively and quantitatively in environmental samples based upon the determination of cyanotoxin synthesis genes is still insufficient. particularly for mc/nod biosynthesis, efforts have been made to correlate cyanotoxin biosynthesis gene abundance with mc/nod occurrence and concentrations (koskenniemi et al., 2007; okello et al., 2010; ostermaier and kurmayer, 2010). in some cases, these efforts did not show the expected correlations and further molecular biological analyses will be needed to understand the relevant influences, for example mobile elements, leading to inactivation (chen et al., 2016). for cyn, stx and antx-a occurrence, the predictability from potential indicative biosynthesis genes is even less explored and considerable further effort will be needed (savela et al., 2015). the occurrence of false positives in laboratory and environmental samples has been reported repeatedly and needs clarification. in particular, for cyn, stx and antx-a the identification of conserved gene (fragments) parts of the core synthesis pathways is still in progress, and needs to identify indicative gene loci to be either used no nco mm er cia l u se on ly j. meriluoto et al.166 singly or in combination. on the other hand, if genes indicative of the above mentioned cyanotoxin synthesis genes are absent, the occurrence of the corresponding cyanotoxins might be excluded. this approach has been partly successful but again depends on the reliability of genetic indication. thus, the idea that molecular tools can guide the application of more sophisticated chemical-analytical techniques through the detection of cyanotoxinindicator genes is still under development. fate, impact and health effects recent reviews on human and animal health aspects include: a health review chapter (codd et al., 2017), an epidemiological assessment of cancers in serbia with reference to cyanobacterial blooms (svirčev et al., 2014a), a review of production, toxinology and toxicology of the antx-a and stx neurotoxins (testai et al., 2016) and a review on epidemiological investigations on health issues related to cyanotoxins (svirčev et al., 2017a). recently reported animal intoxications involving cyanobacterial blooms and cyanotoxins have included dogs in the netherlands, with hantx-a (faassen et al., 2012b) and mc-lr (lürling and faassen, 2013) as likely proximal causes. fish-kills associated with cyanobacterial blooms have continued throughout cyanocost, e.g. in greece and russia (kormas ka, papadimitriou t, chernova e: personal communications). an ongoing likelihood for the role of additional (or alternative) uncharacterised or novel cyanotoxins of health significance was provided in a further, major fish-kill in serbia in which candidate cyanotoxins (cyn, stxs, mcs) were not detected although the c. raciborskii bloom material was lethal by bioassay using artemia salina (svirčev et al., 2016). it is evident that the range of health hazards presented by cyanobacteria and their metabolites is only partially understood. for example, in vitro bioassays have revealed a stimulation of mammalian estrogenic activity by cyanobacterial exudates (sychrova et al., 2012) and a stimulation of proinflammatory responses by mc-lr; the latter effect not being attributable to the long-known inhibition of protein phosphatases by mcs (adamovsky et al., 2015). recognition of the roles of exposure media and routes via which cyanotoxins may present risks to human health has also been increased: one of the first risk assessments of exposure to cyanotoxins from terrestrial sources concerned potential inhalation exposure to airborne mcs from cyanobacterial desert crusts (metcalf et al., 2012). the following paragraphs summarise some examples and case studies which have been well documented in serbia, which is greatly affected by toxic cyanobacterial blooms (simeunović et al., 2010; svirčev et al., 2013; svirčev et al., 2014b). on 14 december 2013, a bloom of planktothrix rubescens (de candolle ex gomont) k. anagnostidis & j. komárek was observed in the vrutci reservoir which is the drinking water source for the city of užice, serbia, where 70,000 inhabitants were potentially exposed to cyanotoxins. the first comprehensive analysis showed the presence of p. rubescens in the treated water at about 10,000 cells per litre, and approximately 1000 cells per litre in a sample from the pipedwater network (institute of public health serbia, 2014). therefore, from 26 december 2013, the use of tap water for drinking and cooking purposes was prohibited by the sanitary inspectorate of the republic of serbia. this initially provoked panic among the citizens. however, the situation was manageable and satisfactorily resolved (at least in the short-term). further analysis showed that the number of p. rubescens cells reached as high as 107,900 cells per millilitre in the reservoir (agency for environmental protection, 2013). the mc-lr concentration in a piped-water network sample was below the who provisional guideline value (1 µg l–1) (institute of public health serbia, 2014). later research (drobac, 2015; svirčev et al., 2017b) confirmed the toxicity of the cyanobacterial biomass by a. salina bioassay. mcs were detected in the lake water by lc-ms/ms (drobac, 2015; svirčev et al., 2017b). mc was also detected in the treated tap water and in fish from the reservoir. additional investigations in the form of questionnaires, epidemiological studies and analysis of cyanotoxins in frozen fish suggested that cyanobacterial blooms might have occurred in the waterbody considerably earlier than the observed bloom in december 2013. furthermore, questionnaire and epidemiological studies indicated the occurrence of gastrointestinal and skin diseases, possible as a result of exposure to cyanobacterial blooms from the reservoir vrutci (đenić, 2014; drobac, 2015; svirčev et al., 2017b).while the vrutci reservoir continued to bloom, a radical but temporary ‘užice solution’ was implemented in order to protect human health, and after 7 february 2014 the abstraction of water from the blooming reservoir was switched to an alternative source, sušičko vrelo. the užice case showed the importance of giving adequate information and advice to the general public. in order to prevent panic and to support the affected population, additional measures should be employed: informative flyers distributed in regions with the problem; daily or weekly updates on the situation; a hotline for the public for further questions and suggestions for dealing with the problem. also, the value of displaying visible, clear and understandable warnings in areas near blooming waters to inform the public of the problem was emphasised, with examples of their successful use in several other countries. the importance of contingency planning, including the prior preparation of warning materials in anticipation of cyanobacterial bloom events was also stressed. no nco mm er cia l u se on ly toxic cyanobacteria in europe research progress through cyanocost 167 prevention and control measures prevention and control measures which could be used to combat toxic cyanobacteria and cyanotoxins have been addressed by some recent european research. the special issue edited by visser et al. (2016b) “cyanobacterial blooms. ecology, prevention, mitigation and control” in the journal aquatic ecology compiles a variety of methods that can be used by water managers to control cyanobacteria in lakes and reservoirs. it provides a state of the art overview on prevention, control and mitigation methods for cyanobacterial blooms. the basis of lake management aiming to reduce cyanobacterial blooms is a comprehensive system analysis which assesses all point and non-point sources of nutrient loading in relation to lake characteristics including morphology, nutrient concentrations and retention times. this multiple approach helps to define the best method to manage cyanobacteria as detailed by stroom and kardinaal (2016). in addition, for an effective implementation of control measures in lakes, these should be preferably developed in collaboration with stakeholders and the citizens in a socio-ecological systems approach (van dolah et al., 2016). the most effective and sustainable way to prevent cyanobacterial blooms is by a strong reduction of the inlake nutrient concentration. this comprises both the reduction of the nutrient load from the catchment (hamilton et al., 2016) and the inflows to the waterbody (fastner et al., 2016). if internal loading is substantial, hypolimnetic aeration (bormans et al., 2016), dredging or sediment capping (douglas et al., 2016) can reduce the release of phosphorus from the sediment. nevertheless, changes in chlorophyll and abundance of cyanobacteria in lakes where phosphorus in the inflow was reduced show that it may take years before the phytoplankton responds, depending largely on the extent of the decrease in the nutrient load achieved (fastner et al., 2016). to overcome cyanobacterial problems in these intervening years, or in cases where sufficient reduction of nutrient loading is too costly or not feasible, control and mitigation measures can be used. the effectiveness of control measures depends on lake characteristics including trophic state, depth and retention time, but also on the cyanobacterial species that bloom in the lake. various cyanobacterial groups have different key functional traits which make them sensitive to certain control measures while being insensitive to others (mantzouki et al., 2016). the requirements and limits of the different methods are summarized by ibelings et al. (2016) and will be briefly listed here. artificial mixing can be a good technique to reduce cyanobacterial blooms in relatively deep systems (>15 m) if the aeration system is well distributed in the lake and strong enough to keep cyanobacterial colonies entrained in the turbulent flow (visser et al., 2016a). cases in which artificial mixing was applied in a way that did not meet these criteria were unsuccessful in reducing cyanobacterial biomass. furthermore, the technique works best for buoyant colony-forming cyanobacteria like microcystis. if mixing is intermittently applied, the technique can also be successful for filamentous cyanobacteria including planktothrix and cylindrospermopsis (antenucci et al., 2005). decreasing the retention time in a lake can reduce the cyanobacterial biomass if the flushing rate is higher than the growth rate of the cyanobacteria and can thus apply for all cyanobacterial species as documented e.g. for microcystis (bowling and baker, 1996; ibelings et al., 1998; verspagen et al., 2006) and dolichospermum (webster et al., 2000; mitrovic et al., 2011). biomanipulation can be successful in lakes where the external and internal loading have been reduced as it can promote the development of submerged macrophytes and help to move from a turbid state to a clear water state (noordhuis et al., 2016; triest et al., 2016). this holds true for water-level management (bakker and hilt, 2016). cyanocides can also be used to reduce the cyanobacterial biomass regardless of the cyanobacterial species. matthijs et al. (2016) discussed the application of conventional and new cyanocides in drinking water reservoirs and bathing waters. especially, hydrogen peroxide appears to be a promising substance for bloom mitigation since cyanobacteria typically show a higher sensitivity to hydrogen peroxide than other groups of aquatic biota. this control agent does not leave any harmful residuals due to its breakdown into oxygen and water, and cyanotoxins degrade rapidly in the treated lakes. there are several further measures that can help in the control of cyanobacteria as discussed by stroom and kardinaal (2016). however, several of the commercially offered endof-pipe techniques should be considered with caution as controlled tests with these techniques were often not very promising (lürling et al., 2016). mitigation techniques including harvesting of scums, shifting the intake-depth in drinking-water reservoirs, deployment of bubble screens or the selective withdrawal of hypolimnetic waters may be used when water quality improvement is urgent for either drinking or recreational use of water. a ban on swimming or bathing may further reduce the risks associated with cyanobacterial bloom exposure at recreational sites. future developments in lake management should aim for clearer guidance in the selection procedure from the suite of available methods, optimization and sustainability of methods and further research into the efficacy and sideeffects of the methods. furthermore, an overview of investments, economic depreciation over time and annual running and maintenance costs will be important elements in the most (cost) effective selection of measures. method selection should also pay attention to the carbon footprint of the different methods (e.g., energy use). there are several conventional and advanced options no nco mm er cia l u se on ly j. meriluoto et al.168 available for the removal/detoxification of cyanotoxins during drinking water treatment (hiskia et al., 2017). physical treatment methods can remove cyanotoxins fully or partly from contaminated water. as their aim is the separation of toxins as intact molecules, further processing for the management of the resulting waste is also required. among these methods, adsorption by activated carbon, either as granular activated carbon (gac) or powdered activated carbon (pac), has been reported to be effective for the removal of several dissolved cyanotoxins. the amount and nature of natural organic matter (nom) in water greatly influences the adsorption and capacity of activated carbon for cyanotoxins. nom is typically present at higher concentrations than cyanotoxins and competes for active sites on the sorbent material. activated carbon technologies may also be combined with biological or membrane technologies to create integrated methods. reverse osmosis and nano-filtration membranes have been tested for mc removal. although efficient, the above technologies should be evaluated in terms of the operational costs and engineering considerations (antoniou et al., 2014). chemical oxidation processes are a promising additional treatment option due to the potential for complete destruction of cyanotoxins, transformation to less toxic by-products and even complete mineralization (sharma et al., 2012). on the other hand, oxidants have been found to lyse cyanobacteria, which may release toxins. it is therefore recommended that coagulation or other separation procedures be applied to remove whole cells and to minimise cyanobacterial cell disruption before oxidation. although studies have been carried out mainly with mcs, all conventional oxidation processes (ozonation, chlorination, oxidation with permanganate) can degrade cyanotoxins in water. among them, ozone seems to be the most effective for mc destruction, reacting preferentially with their double bonds. the ability of ozone to oxidize the adda moiety, which is linked to the toxic properties of mcs, makes treatment with ozone a very promising water detoxification technology for these contaminants. ozone has also been found to be effective for the destruction of antx-a and cyn, although it is less effective with stx (newcombe, 2002). chlorination is also an option for the degradation of mcs and cyn in water, but its efficiency depends on operational parameters, including the nature of the chlorine compounds used. chlorine-based weaker oxidants (and disinfectants) including chlorine dioxide and chloramines have been found to be ineffective for cyanotoxin degradation. contrary to the above, antx-a has been shown to be quite stable upon chlorination (sklenar et al., 2016). permanganate reacts differently with each group of cyanotoxins. it has high reactivity with most mc variants examined and inactivates microcystis aeruginosa cells without significant release of mcs after treatment. the low release of cyanotoxins during treatment in combination with its effectiveness to inactivate m. aeruginosa and oxidize mcs makes permanganate a potential oxidant for their elimination in drinking water treatment (sharma et al., 2012). permanganate is also very reactive with antx-a, with ph playing an important role in the process (ho et al., 2009), while it is not effective with cyn (rodriguez et al., 2007). the reactivity of mcs towards different conventional oxidants is strongly affected by water quality parameters including ph, dissolved organic carbon (doc) and oxidant dose. although there is a general trend for cyanotoxins oxidation (ozone > permanganate > chlorine >>> chlorinebased oxidants), the selection of the appropriate oxidant should be assessed for each particular source of water. advanced oxidation processes (aops) are generally more effective in degrading cyanotoxins than conventional oxidation processes since they generate hydroxyl radicals, which react rapidly with all of the groups of cyanotoxins investigated with high second order reaction rate constants. although the detoxification of water and complete mineralization of cyanotoxins can also be achieved by aops, the roles of water quality parameters (e.g., doc) have to be taken into account. the presence of inorganic and organic water constituents often inhibits the degradation of target pollutants through radical scavenging mechanisms resulting in lower degradation rates. overall, while there are promising results in removing cyanotoxins using aops under laboratory conditions, scale-up studies are needed before these methods are considered as economically feasible and practical sustainable alternatives in water treatment facilities (he et al., 2012; fotiou et al., 2016). in conclusion, effective water treatment strongly depends on both water quality parameters and treatment options. it is recommended that water utilities examine all available alternatives and develop a clear set of goals according to their needs and the expected investment and operational costs before applying measures for cyanotoxin treatment at their facilities. end-user and outreach tools, materials and products while cyanocost has not been the first network focusing on toxic cyanobacteria and cyanotoxins, it is deemed to be the network that has actively involved the highest number of countries as well as reaching the most researchers and other stakeholders. this part of the review concentrates solely on the outreach tools of the cyanocost action as the different approaches taken by the action can serve as an excellent model for similar programmes. when the cyanocost action was planned, it was understood that one of the main objectives of the proposal no nco mm er cia l u se on ly toxic cyanobacteria in europe research progress through cyanocost 169 should be the dissemination of the results of the action throughout the eu and neighbouring countries. the proponents perceived that the cyanotoxin problem in europe was widespread but that different countries and regions had very different management and control mechanisms and policies in place, ranging from extremely low levels of control, to very sophisticated intensive management tools. awareness of the water and health authorities of the problem varied from country to country. the cyanocost action turned out to be a useful vehicle to extend and increase recognition and awareness of the occurrence of cyanobacteria and cyanotoxins, and to make key management tools more widely available to counteract the adverse effects of cyanobacterial mass growths and cyanotoxins. dissemination of the scientific and practical results of the cyanocost action needed to target a broad range of audiences and end-users. these included: i) researchers, especially young ones, from diverse disciplines: analytical chemists, (molecular) biologists, ecologists, toxicologists, water engineers; ii) the drinking water industry, especially managers involved in water quality control; iii) recreational water facility managers; iv) aquaculture and agriculture enterprises and managers; v) environmental laboratories, both public and private; vi) equipment and material designers and manufacturers; vii) public health and clinical professionals, water district authorities and organizations; viii) environmental and public health policy makers at national and european level; ix) non-governmental organizations active in environmental public awareness;(x) regional and national water sports organizations; and xi) the general public. what are the relevant approaches to be considered for a successful dissemination in a scientific project? as exemplified above, cyanocost needed to engage a large number of end-users and stakeholders. the value of different dissemination approaches using databases, webpages, social media, flyers, specialized books, special issues in scientific journals, training schools and shortterm scientific missions (stsms) were considered within the action, and these approaches were developed, based on responses from the action participants and stakeholders. the approaches taken by the action are described in the following text as examples that can be generalised and used in similar projects. a europe-wide rolling database (https://sites.google. com/site/cyanocost/) which was an extension of the 2005 unesco cyanonet database (codd et al., 2005) was developed within the action. the cyanocost database contains useful contacts of researchers, organizations, companies involved in this field, with extended profiles, expertise, publications, projects, collaborations and contact details. links to national risk management policies in use in europe and in neighbouring states, including guidelines and legislation, are included. the database has been originally designed for use by cyanocost participants only but it is expected to be made freely available for the public. in addition to this europe-wide database, a freely-accessible local database, the serbian cyanobacterial database (https://cloud.pmf.uns.ac.rs/index.php/s/v6tervcvcauaxqn), was developed. it is expected that these databases/repositories will be used as sources of information beyond the lifetime of the action. the databases contain material from both pre-cyanocost publications and papers arising from the cyanocost action. based on the cyanocost experience, it is recommended that the following matters receive attention during the lifetime of any action or project establishing public databases: i) intellectual property rights and terms of use; ii) accessibility of information and design of the database; iii) mechanisms to ensure correctness of data; iv) continued updating of the database; and v) maintenance of the database. to address the needs of various professional groups, including water managers, the drinking water industry, the environmental and public health sectors, and academia, cyanocost outcomes were communicated in the form of the three handbooks mentioned in this paper and two special issues in aquatic ecology and advances in oceanography and limnology, i.e. the present themed issue). the handbooks and special issues have been intended to serve as coherent and up-to-date compilations of theoretical and practical knowledge relating to toxic cyanobacteria and the problems they cause. as today the scientific knowledge is scattered over a multitude of scientific individual papers, the value of single-source collections of information is probably appreciated by end-users. furthermore, a decision support tool based on the outcomes of the eu-fp6 project pepcy was translated from german into english and has been made available on-line (https://toxische-cyanobakterien.de/en/). in order to promote networking between various professionals, the action’s management committee meetings have had attendees from the local water and public health authorities, water companies, and the academic community. besides ordinary management committee members, representatives from different institutions in e.g. greece, bulgaria, hungary, spain, the netherlands, etc., participated in the meetings. representatives from other key europe-wide projects such as the netlake cost action or the solutions eu project attended some of the meetings, too. as a reaction to the water crisis in užice, serbia, an international conference ‘water challenges of the future’ was organized in belgrade on 24th march 2014. this scientific and educational meeting was organized by the faculty of sciences of the university of novi sad. eminent experts presented information on cyanobacterial blooms and production of toxic metabolites with the aim of raising preparedness, knowledge and engagement among authorities and citizens. the objective was education of target groups no nco mm er cia l u se on ly j. meriluoto et al.170 including the authorities and citizens of užice, representatives from water-supply facilities, as well as users of recreational waters, and the provision of information on cyanotoxin problems to the institutes of public health, medical and veterinary practitioners, governmental and regional bodies, students of ecology and the general public. furthermore, the merits of introducing guidelines and/or legislation regarding cyanotoxins in serbia (for drinking water, recreational usage and edible fish) were discussed at the meeting with an emphasis on their introduction into national legislation on hazardous substances. the conference provided an example of how an international network can give support to local academics, authorities and other stakeholders in combatting an environmental problem. early-stage researchers (esrs) and other younger professionals were seen as an important group of stakeholders. to increase awareness of cyanobacterial blooms and cyanotoxin problems among young scientists and to promote network-building, summer schools were organized. these included the advanced course on cyanobacteria and cyanotoxins, madrid, june-july 2015, and the 12th summer course on environmental toxicology, brno, june 2015. both summer schools had wide participation from many countries. thematic expert workshops were also organized within the action, including the training school on the use of hydrogen peroxide against cyanobacterial blooms (amsterdam, september 2014) and the analytical workshop on bmaa (wageningen, may 2015). furthermore, a total of 37 stsms were financed to promote mobility between different laboratories in countries and these exchange periods were especially targeted to esrs. the cyanocost webpage (http://www.cyanocost.c om) served as a major forum with information about the action, the subject matter, and as a main communication medium for end-users and others interested in the recognition and risk management of cyanobacterial blooms, cyanotoxins and their impacts. this webpage has received over 30,000 visitors that have visited over 116,000 webpages in 18 months, and the visitors have posted questions, news, announcements, discussions etc. social networks have also been used for dissemination of the cyanocost action. in particular, the twitter account (@cyanocost) was quite successful with over 500 followers and over 2500 ‘tweets’ from june 2013. a cyanocost facebook page has also had over 350 ‘likes’ (followers) and some posts reached to over 1000 followers. taken together, the dissemination efforts of the cyanocost action have allowed an ample distribution of the information regarding cyanobacteria and cyanotoxins not only in cost member countries but also in other countries dealing with cyanobacterial problems such as the united states, australia, etc. gaps in knowledge and actions necessary to strengthen management capabilities the problem of toxic cyanobacterial blooms is primarily linked to non-sustainable consumption by the stillgrowing human population, which brings enormous pressures on global hydrological regimes, nutrient loads and on the global climate. a truly holistic approach will need to be taken at all levels to apply strategies leading to real solutions (paerl et al., 2016) and to develop particular managerial approaches and risk mitigation measures (bullerjahn et al., 2016; ibelings et al., 2016). this section aims to summarize some of the open issues and knowledge gaps which require attention in order to strengthen capabilities for the management of toxic blooms in the future. horizontal and cross-cutting issues with the widening of research on cyanobacteria and their toxins, and the essential involvement of different disciplines, an increasing need is emerging for a proper harmonization and definition of terminology. terms including ‘harmful blue-green algae’, ‘toxic’ or ‘toxigenic’ cyanobacteria, ‘cyanobacteria-l’ and/or ‘algal harmful blooms’, and acronyms such as ‘cyanohabs’ and ’chabs’ are being used in different parts of the world, in different contexts (e.g., marine vs freshwater) and both withinand between different communities (e.g., scientists vs stakeholders). for example, the validity of the terms ‘toxic’ and ‘non-toxic’ as applied to cyanobacteria, depends not only on the presence of cyanotoxins but also on their concentrations and/or the target organisms considered (codd and metcalf, 2014). another emerging problem of terminology relates to current changes in the nomenclature and classification of cyanobacteria (sciuto and moro, 2015; komárek, 2016). current systematics and nomenclature are undergoing rapid changes that can be appreciated only by specialized scientists. however, practitioners and water managers tend to benefit rather from simplification based on mutual agreements among scientists and stakeholders. this approach helps to assure the effective exploitation of different information resources and to streamline the development of management options and policies. the accessibility, reliability, use, and quality of the information are another complex problem related to the risk management of cyanobacteria. for example, regarding spatial information, recent meta-analysis (merel et al., 2013a) identified a disproportionality between well-covered regions including the usa, canada, europe and australia, versus other (highly vulnerable) areas including much of south america, asia or africa (codd et al., 2005, merel et al., 2013a). however, a recent survey has colno nco mm er cia l u se on ly toxic cyanobacteria in europe research progress through cyanocost 171 lated more information on activities in africa (ndlela et al., 2016). further surveys in these continents are needed for truly global mapping of toxic cyanobacteria risks. even existing information may not be fully accessible (merel et al., 2013b). for example, blooms may eventually be monitored and cyanotoxin analyses performed, but the information may be only locally available and not reported to any recognized national or international database(s). thus, the occurrence and potential risks of toxic blooms can be considerably underestimated, and future efforts should aim to implement ‘open environmental data’ principles (e.g., via the geoss system; https://www.earthobservations.org), and to strengthen the role of the public (i.e., via ‘citizen science’) in the field of toxic cyanobacterial blooms. on the contrary, there is a major lag between the speed of newly generated data on cyanobacteria (related to biotechnological advances, next generation sequencing or other ‘omics’; pearson et al., 2016), the actual ‘quality’ of such data (ageing), and very slow progress in the development of bioinformatics selflearning tools that would make use of these ‘big data’. finally, the current ‘publish or perish’ culture in the science research community also negatively affects the field of toxic cyanobacteria. both science and the practical implementation of research results for risk management purposes would greatly benefit from improvements that could be achieved by applying more thorough and critical peerreview during the publication of research on toxic cyanobacteria. surveys surveys of cyanobacterial hazards often combine several methods including: i) taxonomy microscopy determinations of phytoplankton; and/or ii) analyses of cyanotoxins; and/or iii) molecular biology investigations; and/or iv) observations of poisonings or other direct field effects. the information obtained from field surveys may be used not only for immediate managerial actions but is important also for development of models and bloom dynamics predictions, which still remain a major challenge (humbert and fastner, 2017). because of the variability both between and within individual waterbodies, effective predictions are currently possible only a few days in advance, and useful monitoring schemes for risk management require adaptation to local conditions, case-by-case definitions of locations or sampling and analytical details. however, as demonstrated for example by the emls initiative, jointly organized by cost actions cyanocost and netlake, improvements in future monitoring are possible by successful combinations of harmonized methods along with continuous and real-time sampling and standardized analysis methods (humbert and törökné, 2017). recently, numerous multiparameter tools have been developed and integrated into autonomous buoys or drones that may increase our future understanding of detailed spatial and temporal cyanobacterial bloom dynamics. future improvements in bloom monitoring should also consider integration of aerial or space-observations (hunter et al., 2017) and investigation of less explored factors recently suggested as drivers of toxic blooms (carvalho et al., 2011). future surveys of toxigenic cyanobacteria in europe should assess not only traditional cyanotoxins-producers but investigate also alien or invasive species as discussed in previous sections. in particular, distribution and potential toxigenicity of c. bergii, c. ovalisporum or s. aphanizomenoides should not be overlooked. future surveys should also cover picocyanobacteria that often remain overlooked due to their very small size but they have been suggested as potential cyanotoxin-producers by numerous studies (jasser et al., 2017). also benthic cyanobacteria that have previously been associated with animal poisonings still remain to be fully explored (quiblier et al., 2013), including investigations of newly observed couplings in ecology of benthic and pelagic communities (zhang et al., 2015). considering the assessment of individual toxins, modern analytical equipment can easily comply with the most demanding monitoring needs, and robust multitarget methods for simultaneous quantification of multiple compounds are being developed for future implementation. as well, non-target approaches could provide more insights into profiling of cyanotoxins and other metabolites in the future. however, there is a strong need to develop regulatory guidelines for cyanotoxins which would give a higher priority to their monitoring and assessment. another important issue for the future is the current lack of reference materials that are urgently needed for validation and improvements of standardized surveys. as also discussed in previous sections, there are disproportionalities in our knowledge of mcs versus other cyanotoxins. in particular, the occurrence of stx across europe is poorly understood. future surveys should provide insights into stx occurrence and risks specifically focusing e.g. on the association of stx with cylindrospermopsis sp., which has been shown elsewhere but not investigated in europe. with respect to cyn, the main gaps include our understanding of its role in aquatic ecosystems, and studies on factors affecting cyn biosynthesis and biodegradation are needed. also the risks of the neurotoxic amino acid bmaa and its relevance for cyanobacterial blooms will require thorough attention and open discussion among scientists. despite the strengths of molecular biology tools, they are still mainly confined to the field of research, and several issues need to be addressed before they can be widely applied to environmental monitoring. it is accepted that these tools can be developed to provide high-throughput methods that can be used for early-warning purposes durno nco mm er cia l u se on ly j. meriluoto et al.172 ing environmental surveillance. however, due to their high sensitivity, genetic methods in general and pcr assays in particular are prone to errors, i.e. the production of false positive results. for example, in laboratories working routinely on the detection of pathogens by genetic methods, a full catalogue of quality assurance and quality control criteria needs to be met in order to guarantee reliable results. for cyanotoxin synthesis gene detection, in particular, this quality assurance catalogue still needs to be developed; hopefully based on the experience collectively presented in the new handbook on molecular tools (kurmayer et al., 2017). molecular tools can also be expected to make a major contribution to the understanding of the patchy occurrence of the genes for cyanotoxin biosynthesis and of the production of the toxins themselves among cyanobacteria at different levels of taxonomic resolution, e.g. within and between families, genera and species (kurmayer et al., 2015). this variability can be observed for isolated strains and also with environmental samples on a global scale. for example, the biogeography of cyanotoxin biosynthesis might be resolved at a geographically higher resolution, using pcr-based techniques with the aim to map toxigenic genotype occurrence. putting toxigenic genotype distribution on a map could be further combined with phylogenetic analysis to reveal the phylogeographic dependence on cyanotoxin synthesis in nature. effects with respect to effects and impacts, new evidence has been collected throughout the cyanocost action on a range of newly-recognised types of health hazards which cyanobacteria may present. these include for example the production of endocrineor pro-inflammatory-modulators (sychrova et al., 2012; adamovsky et al., 2015), and of retinoic acid receptors which cause malformations during fish embryogenesis (kaya and sano, 2017). several cyanobacterial exposure routes e.g. aerial exposure via inhalation, though long-recognised (codd et al., 1999) require further epidemiological research. in addition, cyanobacterial mats have been discussed as the growth matrix for pathogenic microorganisms (e.g., legionella sp., vibrio cholerae or clostridium sp.; manganelli et al., 2012). future research should critically evaluate the relevance and relative importance of microbial pathogens and further hazards including metals and pesticides to those posed by cyanobacteria. the true incidence of cyanotoxinrelated acute poisonings can hardly be fully quantified but the case reports recently summarized clearly confirm that cyanobacterial blooms present ongoing health risks to both animals and humans (wood, 2016). in addition to traditionally studied liver and neurological symptoms following acute intoxications, new studies are needed to investigate hazards and modes of action (moa) in other tissues or systems (immune-, intestine-, respiratoryetc.). in contrast to the relatively broad knowledge on the acute effects of cyanotoxins, the chronic exposures and related toxic risks and adverse health outcomes remain to be studied in the future (ibelings et al., 2014). preventative and control measures despite the significant progress in our understanding of bloom development and progress in their reduction or prevention, including by in-lake technologies, future improvements are still needed. as pointed out here and summarized elsewhere (ibelings et al., 2016), future lake management would benefit from innovative and clear recommendations on the selection of available methods, their optimization, assessment of efficacy, evaluation of their sustainability and the estimation of potential side effects. furthermore, assessment of the overall cost-effectiveness of individual approaches should be done in the future including evaluation of investments, economic depreciation over time as well as all necessary running and maintenance costs. with regard to the removal of cyanotoxins from waters, current technologies can allow efficient transformations of investigated individual cyanotoxins by at least one known process. however, no universal treatment has been proven to simultaneously and completely remove all cyanotoxins in mixtures, and combinations of different methods (multibarrier concept) should be implemented in advanced treatment plants (merel et al., 2013b). additional measures, such as cyanotoxin removal by home filters, should also be considered under specific situations. improvements are still needed regarding water treatment under variable situations of changing ph or dissolved organic material also with the help of continuous monitoring tools of cyanobacterial breakthroughs (zamyadi et al., 2014). scale-up of promising aops will be needed to evaluate the cost-effectiveness and sustainability of these methods (he et al., 2012; fotiou et al., 2016). finally, the current lack of harmonized international and national regulatory guidelines for cyanotoxins in drinking waters is among the main obstacles for effective risk management in the future, as demonstrated e.g. in the užice case recently (svirčev et al., 2017b). the development of appropriate guidelines and their implementation in most-affected european regions is urgently needed. conclusions as shown by this review, scientific knowledge on toxic cyanobacteria and cyanotoxins has advanced rapidly during recent years. progress has been achieved in e.g. phycological and ecological studies, techniques for cyanotoxin analysis, molecular biology related to toxin biosynthesis, biochemical and medical studies related to cyanotoxin acno nco mm er cia l u se on ly toxic cyanobacteria in europe research progress through cyanocost 173 tions, as well as various management and control aspects. however, the recent cyanotoxin poisonings in europe and elsewhere illustrate clearly that water-users are still experiencing serious cyanobacterial hazards. expertise from various professional groups and collaboration between the stakeholders are necessary in order to help to protect the safety of water consumers and users. answers to some open key questions identified in this review are still pending and will necessitate further research efforts. acknowledgments the authors would like to acknowledge the european cooperation in science and technology, cost action es 1105 ‘cyanocost cyanobacterial blooms and toxins in water resources: occurrence, impacts and management’ for adding value to this paper through networking and knowledge sharing with european experts and researchers in the field. references akcaalan r, köker l, oğuz a, spoof l, meriluoto j, albay m, 2014. first report of cylindrospermopsin production by two cyanobacteria (dolichospermum mendotae and chrysosporum ovalisporum) in lake iznik, turkey. toxins 6:3173-3186. akcaalan r, mazur-marzec h, zalewska a, albay m, 2009. phenotypic and toxicological characterization of toxic nodularia spumigena from a freshwater lake in turkey. harmful algae 8:273-278. adamovsky o, moosova z, pekarova m, basu a, babica p, sindlerova ls, kubala l, blaha l 2015. immunomodulatory potency of microcystin, an important water-polluting cyanobacterial toxin. environ. sci. technol. 49:12457-12464. agency for environmental protection (agencija za zaštitu životne sredine) 2013. accessed on: 21 january 2016. available from: h t t p : / / w w w . s e p a . g o v . r s / i n d e x . p h p ? m e n u =504000000&id=20013&akcija=showarhiva antenucci jp, ghadouani a, burford ma, romero jr, 2005. the long-term effect of artificial destratification on phytoplankton species composition in a subtropical reservoir. freshwater biol. 50:1081-1093. antoniou mg, pelaez ma, song w, o’shea k, ho l, newcombe g, teixeira mr, de la cruz aa, triantis tm, kaloudis t, hiskia a, balasubramanian r, pavagadhi s, han c, sharma v, dixon m, he x, dionysiou dd, 2014. practices that prevent the formation of cyanobacterial blooms in water resources and remove cyanotoxins during physical treatment of drinking water, p. 173-195. in: s. ahuja (ed.), comprehensive water quality and purification. elsevier. b-béres v, gábor v, dobronoki d, gonda s, nagy sa, bácsi i, 2015. effects of cylindrospermopsin producing cyanobacterium and its crude extracts on a benthic green alga competition or allelopathy? mar. drugs 13:6703-6722. bakker es, hilt s, 2016. impact of water level fluctuations on cyanobacterial blooms: options for management. aquat. ecol. 50:485-498. ballot a, fastner j, lentz m, wiedner c. 2010. first report of anatoxin-a-producing cyanobacterium aphanizomenon issatschenkoi in northeastern germany. toxicon 56:964-71. bernard c, ballot a, thomazeau s, maloufi s, furey a, mankiewicz-boczek j, pawlik-skowronska b, capelli c, salmaso n, 2017. appendix 2. cyanobacteria associated with the production of cyanotoxins, p. 503-527. in: j. meriluoto, l. spoof and g.a. codd (eds.), handbook on cyanobacterial monitoring and cyanotoxin analysis. wiley. bláhová l, oravec m, maršálek b, šejnohová l, šimek z, bláha l, 2009. the first occurrence of the cyanobacterial alkaloid toxin cylindrospermopsin in the czech republic as determined by immunochemical and lc/ms methods. toxicon 53:519-524. bohunická m, mareš j, hrouzek p, urajová p, lukeš m, šmarda j, komárek j, gaysina la, sturnecký o, 2015. a combined morphological, ultrastructural, molecular, and biochemical study of the peculiar family gomontiellaceae (oscillatoriales) reveals a new cylindrospermopsin-producing clade of cyanobacteria. j. phycol. 51:1040-1054. bormans m, maršálek b, jančula d, 2016. controlling internal phosphorus loading in lakes by physical methods to reduce cyanobacterial blooms a review. aquat. ecol. 50:407-422. bullerjahn gs, mckay rm, davis tw, baker db, boyer bl, d’anglada lv, doucette gj, ho jc, irwin eg, kling eg, kudela rm, kurmayer r, michalak am, ortiz jd, otten tg, paerl hw, qin bq, sohngen bl, stumpf rp, visser pm, wilhelm sw, 2016. global solutions to regional problems: collecting global expertise to address the problem of harmful cyanobacterial blooms. a lake erie case study. harmful algae 54:223-238. bowling lc, baker pd, 1996. major cyanobacterial bloom in the barwon-darling river, australia, in 1991, and underlying limnological conditions. mar. freshw. res. 47:643-657. bradley wg, mash dc, 2009. beyond guam: the cyanobacteria/bmaa hypothesis of the cause of als and other neurodegenerative diseases. amyotroph. lateral scler. 10:7-20. brient l, lengronne m, bormans m, fastner j, 2009. first occurrence of cylindrospermopsin in freshwater in france. environ. toxicol. 24:415-420. brown nm, mueller rs, shepardson jw, landry zc, morré jt, maier cs, hardy fj, dreher tw, 2016. structural and functional analysis of the finished genome of the recently isolated toxic anabaena sp. wa102. bmc genomics 17:1-18. carvalho l, miller ca, scott em, codd ga, davies ps, tyler an, 2011. cyanobacterial blooms: statistical models describing risk factors for national-scale lake assessment and lake management. sci. total. environ. 409:5353-5358. chen q, christiansen g, deng l, kurmayer r, 2016. emergence of nontoxic mutants as revealed by single filament analysis in bloom-forming cyanobacteria of the genus planktothrix. bmc microbiol. 16:1-12. christiansen g, fastner j, erhard m, börner t, dittmann e, 2003. microcystin biosynthesis in planktothrix: genes, evolution, and manipulation. j. bacteriol. 185:564-572. cirés s, ballot a, 2016. a review of the phylogeny, ecology and toxin production of bloom-forming aphanizomenon spp. and related species within the nostocales (cyanobacteria). harmful algae 54:21-43. cirés s, wörmer l, ballot a, agha r, wiedner c, velázquez d, no nco mm er cia l u se on ly j. meriluoto et al.174 casero mc, queseda a, 2014. phylogeography of cylindrospermopsin and paralytic shellfish toxin-producing nostocales cyanobacteria from mediterranean europe (spain). appl. environ. microbiol. 80:1359-70. codd ga, azevedo smfo, bagchi sn, burch md, carmichael ww, harding wr, kaya k, utkilen hc, 2005. cyanonet a global network for cyanobacterial bloom and toxin risk management. initial situation assessment and recommendations. international hydrobiological programme. technical documents in hydrology no. 76. unesco, paris: 138 pp. codd ga, bell sg, kaya k, ward cj, beattie ka, metcalf js, 1999. cyanobacterial toxins, exposure routes and human health. eur. j. phycol 34:405-415. codd ga, metcalf js, 2014. toxic and non-toxic cyanobacteria: evolving concepts. perspect. phycol. 1: 3-5. codd ga, testai e, funari e, svirčev z (2017). cyanobacteria, cyanotoxins and human health. in: a. hiskia, d. dionysiou, m. antoniou, t. kaloudis, and t. triantis (eds.), water treatment for purification from cyanobacteria and cyanotoxins. wiley (in press). cohen sa, 2012. analytical techniques for the detection of αamino-βmethylaminopropionic acid. analyst 137:1991-2005. cox pa, banack sa, murch sj, rasmussen u, tien g, bidigare rr, metcalf js, morrison lf, codd ga, bergman b, 2005. diverse taxa of cyanobacteria produce β-n-methylamino-lalanine, a neurotoxic amino acid. p. natl. acad. sci. usa 102: 5074-5078. devic e, li d, dauta a, henriksen p, codd ga, marty j-l, fournier d, 2002. detection of anatoxin-a(s) in environmental samples of cyanobacteria using a biosensor with engineered acetylcholinesterases. appl. environ. microbiol. 68:4102-4106. đenić d, 2014. cyanobacterial blooms in vrutci reservoir (cvetanje cijanobakterija u hidroakumulaciji vrutci). master’s thesis, university of novi sad, serbia. đorđević nb, simić sb, ćirić ar, 2015. first identification of the cylindrospermopsin (cyn)-producing cyanobacterium cylindrospermopsis raciborskii (wołoszyńska) seenaya & subba raju in serbia. fresen. environ. bull. 24:3736-3742. douglas gb, hamilton dp, robb ms, pan g, spears bm, lürling m, 2016. guiding principles for the development and application of solid phase phosphorusadsorbents for freshwater ecosystems. aquat. ecol. 50:385-406. drobac d, 2015. human exposure to cyanotoxins and their health effects (putevi izloženosti čoveka cijanotoksinima i njihov uticaj na zdravlje). doctoral dissertation. university of novi sad. duncan mw, 2012. good mass spectrometry and its place in good science. j. mass. spectrom. 47:795-809. dziga d, kokociński m, maksylewicz a, czaja-prokop u, barylski j, 2016. cylindrospermopsin biodegradation abilities of aeromonas sp. isolated from rusałka lake. toxins (basel) 8:55. de la cruz aa, hiskia a, kaloudis t, chernoff n, hill d, antoniou mg, he x, loftin k, o’shea k, zhao c, pelaez m, han c, lynch tj, dionysiou dd, 2013. a review on cylindrospermopsin: the global occurrence, detection, toxicity and degradation of a potent cyanotoxin. environ. sci. process. impacts 15:1979-2003. edwards c, beattie ka, scrimgeour cm, codd ga, 1992. identification of anatoxin-a in benthic cyanobacteria (blue-green algae) and in associated dog poisonings at loch insh, scotland. toxicon 30:1665-1175. elersek t, bláha l, mazur�marzec h, schmidt w, carmeli s, 2017. other cyanobacterial bioactive substances, p. 179-195. in: j. meriluoto, l. spoof and g.a. codd (eds.), handbook on cyanobacterial monitoring and cyanotoxin analysis. wiley. faassen ej, 2014. presence of the neurotoxin bmaa in aquatic ecosystems: what do we really know? toxins 6:1109-1138. faassen ej, antoniou mg, beekman-lukassen w, blahova l, chernova e, christophoridis c, combes a, edwards c, fastner j, harmsen j, hiskia a, ilag ll, kaloudis t, lopicic s, lürling m, mazur-marzec h, meriluoto j, porojan c, viner-mozzini y, zguna n, 2016. a collaborative evaluation of lc-ms/ms based methods for bmaa analysis: soluble bound bmaa found to be an important fraction. mar. drugs 14:45. faassen ej, gillissen f, lürling m, 2012a. a comparative study on three analytical methods for the determination of the neurotoxin bmaa in cyanobacteria. plos one 7:e36667. faassen ej, harkema l, begeman l, lürling m, 2012b. first report of (homo)anatoxin-a and dog neurotoxicosis after ingestion of benthic cyanobacteria in the netherlands. toxicon 60:378-384. fastner j, abella s, litt a, morabito g, vörös l, pálffy k, straile d, kümerlin r, matthews d, phillips mg, chorus i, 2016. combating cyanobacterial proliferation by avoiding or treating inflows with high p load experiences from eight case studies. aquat. ecol. 50:367-384. fastner j, heinze r, humpage ar, mischke u, eaglesham gk, chorus i, 2003. cylindrospermopsin occurrence in two german lakes and preliminary assessment of the toxicity and toxin production of cylindrospermopsis raciborskii cyanobacteria isolates. toxicon 42:313-321. fischbach ma, walsh ct, 2006. assembly-line enzymology for polyketide and nonribosomal peptide antibiotics: logic, machinery, and mechanisms. chem. rev. 106:3468-3496. fotiou t, triantis tm, kaloudis t, o’shea ke, dionysiou dd, hiskia a, 2016. assessment of the roles of reactive oxygen species in the uv and visible light photocatalytic degradation of cyanotoxins and water taste and odor compounds using c-tio2. water res. 90: 52-61. furey a, crowley j, shuilleabhain an, skulberg om, james kj. 2003. the first identification of the rare cyanobacterial toxin, homoanatoxin-a, in ireland. toxicon 41:297-303. gkelis s, zaoutsos n, 2014. cyanotoxin occurrence and potentially toxin producing cyanobacteria in freshwaters of greece: a multi-disciplinary approach. toxicon 78:1-9. gugger m, lenoir s, berger c, ledreux a, druart j-c, humbert j-f, guette c, bernard c, 2005. first report in a river in france of the benthic cyanobacterium phormidium favosum producing anatoxin-a associated with dog neurotoxicosis. toxicon 45:919-28. guzmán-guillén r, prieto ai, moreno im, vasconcelos vm, moyano r, blanco a, cameán am, 2015. cyanobacterium producing cylindrospermopsin causes histopathological changes at environmentally relevant concentrations in subchronically exposed tilapia (oreochromis niloticus). environ. toxicol. 30:261-277. no nco mm er cia l u se on ly toxic cyanobacteria in europe research progress through cyanocost 175 hamilton dp, salmoso n, pearl hw, 2016. catchment control strategies to mitigate harmful cyanobacterial blooms: nitrogen and phosphorus controls. aquat. ecol. 50:351-366. he x, pelaez m, williams c, westrick ja, o’shea ke, hiskia a, triantis t, kaloudis t, de la cruz aa, dionysiou dd, 2012. efficient removal of microcystin-lr by uv-c/h2o2 in synthetic and natural water samples. water res. 46:1501-1510. henriksen p, carmichael ww, an j, moestrup ø, 1997. detection of an anatoxin-a(s)-like anticholinesterase in natural blooms and cultures of cyanobacteria/blue-green algae from danish lakes and in the stomach contents of poisoned birds. toxicon 35: 901-913. hicks lm, moffitt mc, beer ll, moore bs, kelleher nl, 2006. structural characterization of in vitro and in vivo intermediates on the loading module of microcystin synthetase. acs chem. biol. 1:93-102. hiskia a, dionysiou d, antoniou m, kaloudis t, triantis t (eds.), 2017. water treatment for purification from cyanobacteria and cyanotoxins, wiley (in press). ho l, tanis-plant p, kayal n, slyman n, newcombe g, 2009. optimising water treatment practices for the removal of anabaena circinalis and its associated metabolites, geosmin and saxitoxins. j. water health 7:544-56. humbert j-f, fastner j, 2017. ecology of cyanobacteria, p. 1118. in: j. meriluoto, l. spoof, and g.a. codd (eds.), handbook on cyanobacterial monitoring and cyanotoxin analysis, wiley. humbert j-f, törökné a., 2017. new tools for the monitoring of cyanobacteria in freshwater ecosystems, p. 84-88. in: j. meriluoto, l. spoof, and g.a. codd (eds.), handbook on cyanobacterial monitoring and cyanotoxin analysis, wiley. hunter pd, matthews mw, kutser t, tyler an, 2017. remote sensing of cyanobacterial blooms in inland, coastal, and ocean waters, p. 89-99. in: j. meriluoto, l. spoof, and g.a. codd (eds.), handbook on cyanobacterial monitoring and cyanotoxin analysis, wiley. ibelings bw, admiraal w, bijkerk r, ietswaart t, prins h, 1998. monitoring of algae in dutch rivers: does it meet its goals? j. appl. phycol. 2:171-181. ibelings bw, backer lc, kardinaal wea, chorus i, 2014. current approaches to cyanotoxin risk assessment and risk management around the globe. harmful algae 40:63-74. ibelings bw, bormans m, fastner j, visser pm, 2016. cyanocost special issue on cyanobacterial blooms a critical review of the management options for their prevention, control and mitigation. aquat. ecol. 50:595-605. institute of public health serbia “dr milan jovanović batut” 2014. report on the request for access to information of public importance no.8301/1, belgrade. http://uimenaroda.rs/predmeti/li%c4%8dni-stav/190-u%c5%beice-vodosnabdevanje-slika-srbije.html (accessed 21 january 2016). jasser i, callieri c, 2017. picocyanobacteria: the smallest cellsize cyanobacteria, p. 19-27. in: j. meriluoto, l. spoof, and g.a. codd (eds.), handbook on cyanobacterial monitoring and cyanotoxin analysis, wiley. jiang yg, xiao p, yu gl, sano t, pan qq, li rh, 2012. molecular basis and phylogenetic implications of deoxycylindrospermopsin biosynthesis in the cyanobacterium raphidiopsis curvata. appl. environ. microbiol. 78:2256-2263. jiang l, eriksson j, lage s, jonasson s, shams s, mehine m, ilag ll, rasmussen u, 2014a. diatoms: a novel source for the neurotoxin bmaa in aquatic environments. plos one 9(1):e84578. jiang l, kiselova n, rosén j, ilag ll, 2014b. quantification of neurotoxin bmaa (β-n-methylamino-l-alanine) in seafood from swedish markets. sci. rep. 4:6931. jungblut a, neilan b, 2006. molecular identification and evolution of the cyclic peptide hepatotoxins, microcystin and nodularin, synthetase genes in three orders of cyanobacteria. arch. microbiol. 185:107-114. kaštovskỳ j, hauer t, mareš j, krautová m, bešta t, komárek j, desortová b, heteša j, hindáková a, houk v, janeček e, kopp r, marvan p, pumann p, skácelova o, zapomělová e, 2010. a review of the alien and expansive species of freshwater cyanobacteria and algae in the czech republic. biol. invasions 12:3599-3625. kaya k, sano t, 2017. cyanobacterial retinoids, p. 173-178. in: j. meriluoto, l. spoof, and g.a. codd (eds.), handbook on cyanobacterial monitoring and cyanotoxin analysis, wiley. kellmann r, mihali tk, jeon yj, pickford r, pomati f, neilan ba, 2008. biosynthetic intermediate analysis and functional homology reveal a saxitoxin gene cluster in cyanobacteria. appl. environ. microbiol. 74:4044-4053. kiss t, vehovszky a, hiripi l, kovacs a, voros l, 2002. membrane effect of toxins isolated from a cyanobacterium, cylindropermopsis raciborskii, on identified molluscan neurons. comp. biochem. physiol. c toxicol. pharmacol. 131c:167-176. kobos j, błaszczyk a, hohlfeld n, toruńska-sitarz a, krakowiak a, hebel a, sutryk k, grabowska m, toporowska m, kokociński m, messyasz b, rybak a, napiórkowska-krzebietke a, nawrocka l, pełechata a, budzyńska a, zagajewski p, mazur-marzec h, 2013. cyanobacteria and cyanotoxins in polish freshwater bodies. oceanol. hydrobiol. stud. 42:358-378. kokociński m, dziga d, spoof l, stefaniak k, jurczak t, mankiewicz-boczek j, meriluoto j, 2009. first report of the cyanobacterial toxin cylindrospermopsin in the shallow, eutrophic lakes of western poland. chemosphere. 74:669-675. kokociński m, akçaalan r, salmaso n, stoyneva-gärtner mp, sukenik a, 2017. expansion of alien and invasive cyanobacteria, p. 28-39. in: j. meriluoto, l. spoof, and g.a. codd (eds.), handbook on cyanobacterial monitoring and cyanotoxin analysis, wiley. kokociński m, mankiewicz-boczek j, jurczak t, spoof l, meriluoto j, rejmonczyk e, hautala h, vehniäinen m, pawełczyk j, soininen j, 2013. aphanizomenon gracile (nostocales), a cylindrospermopsin-producing cyanobacterium in polish lakes. environ. sci. pollut. res. int. 20:5243-5264. komárek j, 2016. a polyphasic approach for the taxonomy of cyanobacteria: principles and applications. eur. j. phycol. 51:346-353. koreivienė j, kasperovičenė j, 2011. alien cyanobacteria anabaena bergii var. limnetica coutė et preisig from lithuania: some aspects of taxonomy, ecology and distribution. limnologica 41: 325-333. koskenniemi k, lyra c, rajaniemi-wacklin p, jokela j, sivonen k, 2007. quantitative real-time pcr detection of toxic nodularia cyanobacteria in the baltic sea. appl. environ. microbiol. 73:2173-2179. no nco mm er cia l u se on ly j. meriluoto et al.176 kurmayer r, blom jf, deng l, pernthaler j, 2015. integrating phylogeny, geographic niche partitioning and secondary metabolite synthesis in bloom-forming planktothrix. isme j. 9:909-921. kurmayer r, sivonen k, wilmotte a, salmaso n (eds.), 2017. molecular tools for the detection and quantification of toxigenic cyanobacteria. wiley (in press). lage s, burian a, rasmussen u, costa pr, annadotter h, godhe a, rydberg s, 2015. bmaa extraction of cyanobacteria samples: which method to choose? environ. sci. pollut. res. int. 23(1):338-350. ledreux a, thomazeau s, catherine a, duval c, yéprémian c, marie a, et al. 2010. evidence for saxitoxins production by the cyanobacterium aphanizomenon gracile in a french recreational water body. harmful algae. 10:88-97. li r, carmichael ww, brittain je, eaglesham gk, shaw gr, mahakhant a, noparatnaraporn n, yongmanitchai w, kaya k, watanabe mm, 2001. isolation and identification of the cyano-toxin cylindrospermopsin and deoxycylindrospermopsin from a thailand strain of cylindrospermopsis raciborskii (cyanobacteria). toxicon 39:973-980 lürling m, faassen ej, 2013. dog poisoning associated with a microcystis aeruginosa bloom in the netherlands. toxins 5:556-567. lürling m, waajen g, de senerpont domis ln, 2016. evaluation of several end-of-pipe measures proposed to control cyanobacteria. aquat. ecol. 50:499-520. manganelli m, scardala s, stefanelli m, palazzo f, funari e, vichi s, buratti fm, testai e, 2012. emerging health issues of cyanobacterial blooms. ann. ist. super. sanita 48:415-428. manti g, mattei d, messineo v, melchiorre s, bogialli s, sechi n, casiddu p, luglie’ a, di brizio m, bruno m, 2005. first report of cylindrospermopsis raciborskii in italy. harmful algae news 28: 8-9. mantzouki e, visser pm, bormans m, ibelings bw, 2016. understanding the key ecological traits of cyanobacteria as a basis for their management and control in changing lakes. aquat. ecol. 50:333-350. matthijs hcp, jančula d, visser pm, maršálek b, 2016. existing and emerging cyanocidal compounds, new perspectives for cyanobacterial bloom mitigation. aquat. ecol. 50:443-460. mazmouz r, chapuis-hugon f, pichon v, méjean a, ploux o, 2011. the last step of the biosynthesis of the cyanotoxins cylindrospermopsin and 7-epi-cylindrospermopsin is catalysed by cyri, a 2-oxoglutarate-dependent iron oxygenase. chembiochem 12:858-862. mazur-marzec h, kaczkowska mj, blaszczyk a, akcaalan r, spoof l, meriluoto j, 2013. diversity of peptides produced by nodularia spumigena from various geographical regions. mar. drugs 11:1-19. méjean a, mann s, maldiney t, vassiliadis g, lequin o, ploux o, 2009. evidence that biosynthesis of the neurotoxic alkaloids anatoxin-a and homoanatoxin-a in the cyanobacterium oscillatoria pcc 6506 occurs on a modular polyketide synthase initiated by l-proline. j. am. chem. soc. 131:7512-7513. méjean a., paci g, gautier v, ploux o, 2014. biosynthesis of anatoxin-a and analogues (anatoxins) in cyanobacteria. toxicon 91:15-22. merel s, villarin mc, chung k., snyder s, 2013a spatial and thematic distribution of research on cyanotoxins. toxicon 76:118-131. merel s, walker d, chicana r, snyder s, baures e, thomas o, 2013b state of knowledge and concerns on cyanobacterial blooms and cyanotoxins. environ. int.. 59:303-327. meriluoto j, spoof l, codd ga (eds.), 2017. handbook of cyanobacterial monitoring and cyanotoxin analysis. wiley. messineo v, bogialli s, melchiorre s, sechi n, lugliè a, casiddu p, mariani ma, padedda bm, di corcia a, mazza r, carloni e, bruno m, 2009. cyanobacterial toxins in italian freshwaters. limnologica 39:95-106. metcalf js, richer r, cox pa, codd ga, 2012. cyanotoxins in desert environments may present a risk to human health. sci. total environ. 421-422:118-123. mihali tk, kellmann r, muenchhoff j, barrow kd, neilan ba, 2008. characterization of the gene cluster responsible for cylindrospermopsin biosynthesis. appl. environ. microbiol. 74:716-722. mitrovic sm, hardwick l, dorani f, 2011. use of flow management to mitigate cyanobacterial blooms in the lower darling river, australia. j. plankton res. 2:229-241. moffitt mc, neilan ba, 2004. characterization of the nodularin synthetase gene cluster and proposed theory of the evolution of cyanobacterial hepatotoxins. appl. environ. microbiol. 70:6353-6362. murch sj, cox pa, banack sa, 2004. a mechanism for slow release of biomagnified cyanobacterial neurotoxins and neurodegenerative disease in guam. proc. natl. acad. sci. u.s.a. 101: 12228-12231. murray sa, mihali tk, neilan ba, 2011. extraordinary conservation, gene loss, and positive selection in the evolution of an ancient neurotoxin. mol. biol. evol. 28(3):1173-1182. ndlela ll, oberholster pj, van wyk jh, cheng ph, 2016. an overview of cyanobacterial bloom occurrences and research in africa over the last decade. harmful algae 60:11-26 newcombe g, 2002. removal of algal toxins from drinking water using ozone and gac. awwa research foundation and american water works association, denver, co, usa: 133 pp. noordhuis r, van zuidam bg, peeters ethm, 2016. further improvements in water quality of the dutch borderlakes: two types of clear states at different nutrient levels. aquat. ecol. 50:521-540. ohtani i, moore re, runnegar mtc, 1992. cylindrospermopsin: a potent hepatotoxin from the blue-green alga cylindrospermopsis raciborskii. j. am. chem. soc. 114:7941-7942 okello w, ostermaier v, portmann c, gademann k, kurmayer r, 2010. spatial isolation favours the divergence in microcystin net production by microcystis in ugandan freshwater lakes. water res. 44:2803-2814. o’neil jm, davis tw, burford ma, gobler cj, 2012. the rise of harmful cyanobacteria blooms: the potential roles of eutrophication and climate change. harmful algae 14:313334. onodera h, oshima y, henriksen p, yasumoto t, 1997. confirmation of anatoxin-a(s), in the cyanobacterium anabaena lemmermannii, as the cause of bird kills in danish lakes. toxicon 35: 1645-1648. osswald j, rellán s, gago-martinez a, vasconcelos v, 2009. production of anatoxin-a by cyanobacterial strains isolated no nco mm er cia l u se on ly toxic cyanobacteria in europe research progress through cyanocost 177 from portuguese fresh water systems. ecotoxicology 18:1110-1115. ostermaier v, kurmayer r, 2010. application of real-time pcr to estimate toxin production by the cyanobacterium planktothrix sp. appl. environ. microbiol. 76:3495-3502. paerl hw, gardner ws, havens ke, joyner ar, mccarthy mj, newell se, qin bq, scott jt, 2016. mitigating cyanobacterial harmful algal blooms in aquatic ecosystems impacted by climate change and anthropogenic nutrients. harmful algae 54:213-222. pearson la, dittmann e, mazmouz r, ongley se, p. d’agostino m, neilan ba, 2016. the genetics, biosynthesis and regulation of toxic specialized metabolites of cyanobacteria. harmful algae 54:98-111. poniedziałek b, rzymski p, wiktorowicz k, 2014. toxicity of cylindrospermopsin in human lymphocytes: proliferation, viability and cell cycle studies. toxicol. in vitro 5: 968-974. poniedziałek b, rzymski p, kokociński m, karczewski j, 2015. toxic potencies of metabolite(s) of non-cylindrospermopsin producing cylindrospermopsis raciborskii isolated from temperate zone in human white cells. chemosphere 120:608-614. quesada a, moreno e, carrasco d, paniagua t, wormer l, de hoyos c, sukenik a, 2006. toxicity of aphanizomenon ovalisporum (cyanobacteria) in spanish water reservoir. eur. j. phycol. 41:39-45. quiblier c, wood s, echenique-subiabre i, heath m, villeneuve a, humbert jf, 2013. a review of current knowledge on toxic benthic freshwater cyanobacteria ecology, toxin production and risk management. water res. 47:5464-5479. rantala a, fewer dp, hisbergues m, rouhiainen l, vaitomaa j, börner t, sivonen k, 2004. phylogenetic evidence for the early evolution of microcystin synthesis. proc. natl. acad. sci. u.s.a. 101:568-573. rantala-ylinen a, kana s, wang h, rouhiainen l, wahlsten m, rizzi e, berg k, gugger m, sivonen k, 2011. anatoxin-a synthetase gene cluster of the cyanobacterium anabaena sp strain 37 and molecular methods to detect potential producers. appl. environ. microbiol. 77:7271-7278. rapala j, robertson a, negri ap, berg ka, tuomi p, lyra c, erkomaa k, lahti k, hoppu k, lepistö l, 2005. first report of saxitoxin in finnish lakes and possible associated effects on human health. environ. toxicol. 20:331-40. réveillon d, abadie e, séchet v, masseret e, hess p, amzil z, 2015. β-n-methylamino-l-alanine (bmaa) and isomers: distribution in different food web compartments of thau lagoon, french mediterranean sea. mar. environ. res. 110:8-18. réveillon d, séchet v, hess p, amzil z, 2016a. production of bmaa and dab by diatoms (phaeodactylum tricornutum, chaetoceros sp., chaetoceros calcitrans and, thalassiosira pseudonana) and bacteria isolated from a diatom culture. harmful algae 58:45-50. réveillon d, séchet v, hess p, amzil z, 2016b. systematic detection of bmaa (β-n-methylamino-l-alanine) and dab (2,4-diaminobutyric acid) in mollusks collected in shellfish production areas along the french coasts. toxicon 110:35-46. rodriguez e, sordo a, metcalf js, acero jl, 2007. kinetics of the oxidation of cylindrospermopsin and anatoxin-a with chlorine, monochloramine and permanganate. water res. 41:2048-2056. rzymski p, poniedziałek b, 2014. in search of environmental role of cylindrospermopsin: a review on global distribution and ecology of its producers. water res. 66:320-337. rosén j, westerberg e, schmiedt s, hellenäs ke, 2016. bmaa detected as neither free nor protein bound amino acid in blue mussels. toxicon 109:45-50. sahindokuyucu kocasari f, gulle i, kocasari s, pekkaya s, mor f, 2015. the occurrence and levels of cyanotoxin nodularin from nodularia spumigena in the alkaline and salty lake burdur, turkey. j. limnol. 74:530-536. saker ml, neilan ba, 2001. varied diazotrophies, morphologies, and toxicities of genetically similar isolates of cylindrospermopsis raciborskii (nostocales, cyanophyceae) from northern australia. appl. environ. microbiol. 67:1839-1845. savela h, spoof l, perälä n, preede m, lamminmäki u, nybom s, häggqvist k, meriluoto j, vehniäinen m, 2015. detection of cyanobacterial sxt genes and paralytic shellfish toxins in freshwater lakes and brackish waters on åland islands, finland. harmful algae 46:1-10. sciuto k, moro i, 2015. cyanobacteria: the bright and dark sides of a charming group. biodivers. conserv. 24:711-738. shalev-alon g, sukenik a, livnah o, schwarz r, kaplan a, 2002. a novel gene encoding amidinotransferase in the cylindrospermopsin-producing cyanobacterium aphanizomenon ovalisporum. fems microbiol. lett. 209:87-91. sharma vk, triantis tm, antoniou mg, he x, pelaez m, han c, song w, o’shea ke, de la cruz aa, kaloudis t, hiskia a, dionysiou dd, 2012. destruction of microcystins by conventional and advanced oxidation processes: a review. sep. purif. technol. 91:3-17. sierosławska a, rymuszka a, 2015. effects of cylindrospermopsin on a common carp leucocyte cell line. j. appl. toxicol. 35:83-89. simeunović j, svirčev z, karaman m, knezević p, melar m 2010. cyanobacterial blooms and first observation of microcystin occurrences in freshwater ecosystems in vojvodina region (serbia). fresen. environ. bull. 19:198-207. sklenar k, westrick j, szlag d, 2016. managing cyanotoxins in drinking water: a technical guidance manual for drinking water professionals. american water works association and water research foundation, denver, co, usa: 61 pp. spoof l, berg ka, rapala j, lahti k, lepistö l, metcalf js, codd ga, meriluoto j, 2006. first observation of cylindrospermopsin in anabaena lapponica isolated from the boreal environment (finland). environ. toxicol. 21:552-560. stirling dj, quilliam ma, 2001. first report of the cyanobacterial toxin cylindrospermopsin in new zealand. toxicon 39:1219-1222. stroom jm, kardinaal wea, 2016. how to combat cyanobacterial blooms: strategy toward preventive lake restoration and reactive control measures. aquat. ecol. 50:521-576. sukenik a, quesada a, salmaso n, 2015. global expansion of toxic and non-toxic cyanobacteria: effect on ecosystem functioning. biodivers. conserv. 24:889-908. svirčev z, drobac d, tokodi n, lužanin z, munjas am, nikolin b, vuleta d, meriluoto j, 2014a. epidemiology of cancers in serbia and possible connection with cyanobacterial blooms. j. env. sci. health c environ. carcinog. ecotoxicol. rev. 32:319-337. svirčev z, drobac d, tokodi n, codd ga, meriluoto j, 2017a. no nco mm er cia l u se on ly j. meriluoto et al.178 toxicology of microcystins with reference to cases of human intoxications and epidemiological investigations of exposures to cyanobacteria and cyanotoxins. arch. toxicol. 91:621-650. svirčev z, drobac d, tokodi n, đenić d, simeunović j, hiskia a, kaloudis t, mijović b, šušak s, protić m, vidović m, onjia a, nybom s, važić t, palanački malešević t, dulić t, pantelić d, vukašinović m, meriluoto j, 2017b. lessons from the užice case: how to complement analytical data, p. 298-308. in: j meriluoto, l spoof, and ga codd (eds.), handbook of cyanobacterial monitoring and cyanotoxin analysis, wiley. svirčev z, obradović v, codd ga, marjanović p, spoof l, drobac d, tokodi n, petković a, nenin t, simeunović j, važić t, meriluoto m, 2016. massive fish mortality and cylindrospermopsis raciborskii bloom in aleksandrovac lake. ecotoxicology 25:1352-1363. svirčev z, simeunović j, subakov-simić g, krstić s, pantelić d, dulić t, 2013. cyanobacterial blooms and their toxicity in vojvodina lakes, serbia. int. j. environ. res. 7:745-758. svirčev z, tokodi n, drobac d, codd ga, 2014b. cyanobacteria in aquatic ecosystems in serbia: effects on water quality, human health and biodiversity. system. biodivers. 12:261-270. sychrova e, stepankova t, novakova k, blaha l, giesy jp, hilscherova h, 2012. estrogenic activity in extracts and exudates of cyanobacteria and green algae. environ. int. 39:134-140. testai e, scardala s, victu s, buratti m, funari e, 2016. risk to human health associated with the environmental occurrence of cyanobacterial neurotoxic alkaloids anatoxins and saxitoxins. crit. rev. toxicol. 46:285-419. tillett d, dittmann e, erhard m, vondöhren h, börner t, neilan ba, 2000. structural organization of microcystin biosynthesis in microcystis aeruginosa pcc7806: an integrated peptide-polyketide synthetase system. chem. biol. 7:753-764. triest l, stiers i, van onsem s, 2016. biomanipulation as a nature-based solution to reduce cyanobacterial blooms. aquat. ecol. 50:461-484. van dolah er, paolisso m, sellner k, place a, 2016. employing a socio-ecological systems approach to engage harmful algal blooms stakeholders. aquat. ecol. 50:577-594. verspagen jm, passarge j, jöhnk kd, visser pm, peperzak l, boers p, laanbroek hj, huisman j, 2006. water management strategies against toxic microcystis blooms in the dutch delta. ecol. appl. 16:313-327. visser pm, ibelings bw, bormans m, huisman j, 2016a. artificial mixing to control cyanobacterial blooms: a review. aquat. ecol. 50:423-442. visser pm, ibelings bw, fastner j, bormans m (eds.), 2016b. special issue ‘cyanobacterial blooms. ecology, prevention, mitigation and control.’ aquat ecol. 50:327-605. webster it, sherman bs, bormans, m, jones g, 2000. management strategies for cyanobacterial blooms in an impounded lowland river. regul. rivers res. manag. 16:513-525. wimmer km, strangman wk, wright jlc, 2014. 7-deoxydesulfo-cylindrospermopsin and 7-deoxy-desulfo-12-acetylcylindrospermopsin: two new cylindrospermopsin analogs isolated from a thai strain of cylindrospermopsis raciborskii. harmful algae 37: 203-206. wood r, 2016. acute animal and human poisonings from cyanotoxin exposure a review of the literature. environ. int. 91:276-282. zamyadi a, dorner s, ndong m, ellis d, bolduc a, bastien c, prevost m, 2014. application of in vivo measurements for the management of cyanobacteria breakthrough into drinking water treatment plants. environ. sci. process. impacts 16:313-323. zervou s-k, christophoridis c, kaloudis t, triantis tm, hiskia a, 2017. new spe-lc-ms/ms method for simultaneous determination of multi-class cyanobacterial and algal toxins. j. hazard. mater. 323:56-66. zhang qt, warwick rm, mcneill cl, widdicombe ce, sheehan a, widdicombe s, 2015. an unusually large phytoplankton spring bloom drives rapid changes in benthic diversity and ecosystem function. progr. oceanogr. 137:533-545. no nco mm er cia l u se on ly layout 1 introduction cyanobacteria are ubiquitous microorganisms, which cause environmental and health issues in lakes and reservoirs. the massive development of cyanobacteria alters the bio-chemical equilibrium in the water basin, possibly resulting in a deterioration of the water quality. environmental alterations caused by climate changes, eutrophication and hydrological changes are considered the major factors favoring the dominance and the spreading of cyanobacteria in freshwaters (schindler, 2006; paerl and huisman, 2009; schopf, 2012; hamilton et al., 2016). many blooms-forming cyanobacteria have the potential of producing toxic secondary metabolites, which cause serious illnesses in case of human exposure (by ingestion, inhalation or skin contact). for this reason, many countries have issued specific regulations for preventing human exposure to cyanobacterial toxins both from drinking and recreational activities (welker and von döhren, 2006; iarc, 2010; chorus, 2012). most of the bioactive metabolites produced by cyanobacteria can be classified in two major chemical classes: peptides and alkaloids. microcystins (mcs) and nodularins (nods) belong to the former class with mcs representing the most common toxins. anatoxins (atxs), cylindrospermopsins (cyns) and paralytic shellfish poisons (psps) represent instead the most common toxic alkaloids found in cyanobacteria. the just cited compounds have attracted a lot of interest by the research community which has led to the development of efficient analytical procedures, and standardized extraction and analysis protocols are already available or on the way to be so (codd, 1995; zurawell et al., 2005; van apeldoorn et al., 2007). other bioactive compounds, sometimes produced by cyanobacteria in comparable amounts with the toxins reported above, have gained less interest and, consequently, there is a very limited knowledge about their occurrence. for example, anabaenopeptins, aeruginosins, microginins and microviridins are metabolites that are potentially toxic for mammals (shin et al., 1995; neumann et al., 1997; advances in oceanography and limnology, 2017; 8(1): 22-32 article doi: 10.4081/aiol.2017.6381 this work is licensed under a creative commons attribution-noncommercial 4.0 international license (cc by-nc 4.0). a comparative study of the metabolic profiles of common nuisance cyanobacteria in southern perialpine lakes leonardo cerasino,1* camilla capelli,1,2 nico salmaso1 1department of sustainable agro-ecosystems and bioresources, iasma research and innovation centre, fondazione edmund mach (fem), via e. mach 1, 38010 san michele all’adige (tn); 2department of biology, university of florence, via madonna del piano, 6, 50019 sesto fiorentino (fi), italy *corresponding author: leonardo.cerasino@fmach.it abstract this work allowed the comparison of the metabolic profiles of the most important cyanobacteria species in southern perialpine lakes, namely aphanizomenon flos-aquae, dolichospermum lemmermannii, microcystis aeruginosa, planktothrix rubescens, and tychonema bourrellyi. monospecific cultures were obtained from samples of 3 different natural lakes (garda, idro, and caldonazzo). lcms/ms analyses were conducted on strains. a first set of experiments was aimed at assessing the presence of the best known toxins (microcystins, nodularins, (homo)anatoxin-a, cylindrospermopsins, paralytic shellfish poisons) in the cultures. results of this screening study revealed that m. aeruginosa and p. rubescens produced toxic peptides (microcystins), t. bourrellyi produced toxic alkaloids (anatoxin-a and possibly some paralytic shellfish toxins), aph. flos-aquae and d. lemmermannii did not produce any of the analyzed toxins. m. aeruginosa and p. rubescens showed typical microcystin production with lr form dominant in the former, and rrdm form dominant in the latter. a second set of experiments was aimed at comparing the capability of the 5 cyanobacterial species to produce peptidic secondary metabolites. for this purpose, an untargeted peptidomic analysis was conducted on the strains. the analysis allowed revealing globally 328 metabolites, spanning in a mass range between 400 and 2000 da. the majority of compounds with masses in the 5001200 da range (corresponding to the majority of peptidic secondary metabolites) resulted to be produced by m. aeruginosa and p. rubescens strains, thus indicating a higher ability of these species to produce non-ribosomal peptides compared to the others. 27 metabolites out of 328 could be putatively assigned to specific classes of compounds: microcystins, aeruginosins and anabaenopeptins were the most represented classes of compounds, and were mostly found in m. aeruginosa and p. rubescens strains. key words: cyanobacteria; perialpine lakes; metabolic profiles; cyanotoxins; lc-ms. received: november 2016. accepted: may 2017. no nco mm er cia l u se on ly metabolic profiles of cyanobacteria in perialpine lakes 23 murakami et al., 2000; blom et al., 2006; bubik et al., 2008; ersmak et al., 2008). since they can contribute to the total toxicity of a given cyanobacteria population, the lack of information about their presence can potentially bias a correct risk assessment. besides the health issue aspects, some of these metabolites have attracted interest for biotechnological applications, for example as source of new therapeutic agents (sivonen and borner, 2008; anas and harada, 2016). genetic diversity in cyanobacteria is reflected also in the secondary metabolites profiles. different species exhibit different metabolic profiles, but also different strains of the same species can show substantial differences in the metabolites profile (moore, 1996; janse et al., 2005; kardinaal et al., 2007; yepremian et al., 2007; rohrlack et al., 2008; agha et al., 2014). in addition, genetically identical organisms can exist as different chemotypes. this might be due to the high plasticity of the metabolic pathways, which can re-arrange with no or minimal genetic changes (fischbach et al., 2008; nikolouli and mossialos, 2012). lakes in the southern perialpine region are experiencing a change in their cyanobacteria composition and structure. five toxic species are commonly found in these lakes: aphanizomenon flos-aquae ralfs ex bornet & flahault, dolichospermum lemmermannii (richter) p.wacklin, l.hoffmann & j.komárek, microcystis aeruginosa (kützing) kützing, planktothrix rubescens (de candolle ex gomont) anagnostidis & komárek, and tychonema bourrellyi (j.w.g.lund) anagnostidis & komárek. t. bourellyi has appeared in recent years and is replacing other species (salmaso et al., 2016). changes in cyanobacteria populations lead to changes in toxic potential, as different species have a very different toxin diversity. these changes have therefore a great impact on the management of the risk connected with cyanobacterial blooms. t. bourrellyi for example is an atx producer that is becoming dominant in lakes formerly dominated by mc producers (salmaso et al., 2016). in order to have a clear picture of the toxic potential of these five species, we have conducted a detailed investigation aimed at identification of the toxins produced by them. we used sensitive and selective lc-ms/ms techniques for the analysis of a broad spectrum of known toxins (mcs, nods, atxs, cyns, and psps). in order to get more information on the secondary metabolites of these species and possibly identify additional bioactive ones, we analyzed the peptidomic profiles of the five species. methods isolation of strains and culturing methods samples were collected from three lakes located in the eastern italian perialpine area: lakes garda, idro and caldonazzo. strains of p. rubescens, t. bourrellyi, and d. lemmermannii were isolated from lake garda in summer/early autumn of 2014. strains of aph. flos-aquae and m. aeruginosa were isolated, respectively, in april 2014 from lake idro and in august 2015 from lake caldonazzo. water samples were collected in the deepest point of the basins, by vertical tows from 30 m to the surface with 25 cm diameter plankton net (80 µm mesh) and maintained at 20°c until processed (within 24 h). single strains were isolated under a macroscope (wild m420) using a microcapillary and identified according to komárek and anagnostidis (2005). after 3-4 washings in z8 medium (kotai, 1972), strains were grown, under nonaxenic conditions, in microtiter plates filled with 3 ml z8 medium, and then transferred to the final volume of 150 ml z8 medium in cell culture flasks (cellstar, greiner bio-one gmbh, kremsmünster, austria). aph. flos-aquae was grown in z8 medium without nitrogen (kotai, 1972). all cultures were grown in climatic simulation chamber (proclimatic, imola, bo, italy) at 20°c, using continuous light (25 µmol m–2 s–1) in the case of p. rubescens, t. bourrellyi, and aph. flos-aquae, and a 16:8 h light:dark photoperiod in the case of d. lemmermannii and m. aeruginosa. culturing conditions were those allowing an optimal growth for each cyanobacterial species. metabolites extraction and lc-ms/ms analysis all solvents and reagents used in the following procedures were of lc-ms grade and where all provided by sigma-aldrich (milan, italy), if not otherwise specified. lc-ms/ms analysis were performed using a waters acquity uplc system, directly coupled to a sciex 4000 qtrap hybrid mass spectrometer equipped with a turbo ion spray interface. biomass was harvested from the culture by filtration (between 200 and 250 ml) on 1.2 µm gf/c filters (whatman-ge healthcare life sciences, little chalfont, uk). extraction of metabolites was achieved by extraction with 6 ml of acetonitrile/water mixture (60/40 v/v) containing 0.1% formic acid; the sample was firstly homogenized (omni th probe homogenizer, omni-inc., kennesaw, ga, usa) for 5 min and then sonicated (omniruptor4000 probe sonicator, omni-inc.) for 4 min, using 160w power in pulsed mode (50%). the solution was separated from the pellet by centrifugation (eba 20, hettich, tuttlingen, germany) for 6 min at 9850 g. the pellet was then treated with a second aliquot of extraction mixture (6 ml) and sonicated. after centrifugation, the two aliquots were put together and concentrated in a centrifugal evaporator (mivacduo, genevac ltd., ipswich, uk) down to approximately 2 ml. acetonitrile was then added up to reach a 60/40 ratio between organic solvent and water. the solution was then filtered on 0.2 µm pore size rc syringe filters (phenomenex, castel maggiore, bo, italy), and analyzed by lc-ms/ms. no nco mm er cia l u se on ly l. cerasino et al.24 targeted mcs and nods analysis was performed using reverse phase chromatography, using a phenomenex kinetex xb-c18 column (1.7 μm particle size, 2.1×50 mm). the mass detector was operated in the positive electro spray mode (esi+) using the multiple reaction monitoring (mrm) scanning mode. the method was optimized for the detection of commercially available analytical standards mc-rr, [d-asp3]-rr, yr, lr, [dasp3]-lr, wr, la, ly, lw, lf, nod-r (sigma-aldrich co., st. louis, mo, usa) and their variants (demethylated forms), and could potentially detect up to 40 mc congeners. a detailed description of the method is reported in cerasino et al. (2016). targeted alkaloids analysis was performed by hilic, using an ascentis express oh5 column (2.7 µm particle size, 50x2.1 mm, sigmaaldrich co., usa). elution was achieved by a binary gradient of eluents a (1% acetonitrile in water, containing 10 mm ammonium formate and 10 mm formic acid) and b (95% acetonitrile in water, containing 10 mm ammonium formate and 10 mm formic acid) according to the following scheme: t=0 (90% b), t=5 min (50% b), t=7 min (90% b), t=8 min (90% b). flow rate was 0.3 ml min–1 and total run time was 8 min. standard injection volume was 2 µl. mass detector was operated in scheduled mrm mode using positive electrospray ionization (esi+). general settings were as follows: ion spray voltage 5000 v, entrance potential 10 v, cell exit potential 10 v, and interface heater temperature 300°c. for toxin identification, two transitions were monitored for each analyte (tab. 1). the method was set up using certified analytical standards: homoatx (novakits, france), atx (tocris, uk), cyn (vinci biochem, italy), saxitoxin (stx), decarbamoyl stx (dcstx), neostx, gonyautoxin 1 (gtx 1), gtx4, gtx5, c1 and c2, (nrc-cnrc, canada). a reference chromatogram is provided in fig. 1. retention times of analytes showed excellent stability over time. the method was suitable for the quantification of these toxins, at least in the working range between 0.2 and 200 µgl–1. nevertheless, only a semi-quantitative analysis was performed in this investigation owing to the heterogeneity of the cultures. a tentative detection of other psp congeners not available as pure standards was performed, using transitions and mssettings (cf. tab. 1) taken from relevant literature (dell’aversano et al., 2005; halme et al., 2012). untargeted metabolic profiling experiments were contab. 1. chromatographic and mass spectrometric parameters for alkaloids’ targeted analysis. [m+h]+ indicates the observed “quasimolecular” ion for any given toxin. it usually corresponds to the precursor ion in mrm transitions. in some cases, however, other ions are used as precursor because the intensity of the [m+h]+ adducts is too low (c1/2 and c3/4 toxins). for each compound two mrm transitions have been monitored. in the case of the pairs of isomeric compounds gtx1/4 and c1/2, the two transitions of the pairs coincide. limits of quantification (loq) are also reported for each compound. toxin rt (min) [m+h]+ precursor ion (m/z) product ions (m/z) dp (v) ce (v) loq (µgl–1) hatx 0.82 180 180 145 60 23 2.0 135 60 23 atx 0.98 166 166 149 60 21 0.5 131 60 24 cyn 2.50 416 416 194 80 53 2.0 336 80 32 gtx1/4 3.48 (gtx1)+3.72 (gtx4) 412 412 332 40 20 15.0 412 332 ([m+h-so3]+) 314 86 27 c1/2 3.66 (c1)+3.93 (c2) 476 493 ([m+h+nh3]+) 396 40 13 10.0 476 396 ([m+h-so3]+) 298 80 27 neostx 4.08 316 316 298 70 25 30.0 220 70 25 gtx5 4.10 380 380 300 50 21 5.0 282 50 25 stx 4.10 300 300 204 80 35 10.0 138 80 40 dcstx 4.17 257 257 239 80 25 20.0 126 80 29 dcneostx* 273 273 255, 225, 179 70 25 gtx2/3* 396 396 378, 316 70 25 dcgtx2/3* 353 353 335, 273 70 25 gtx5* 380 380 300, 282 70 25 c3/4* 492 412 ([m+h-so3]+) 332, 314, 138 70 25 *compounds which are tentatively analyzed. no nco mm er cia l u se on ly metabolic profiles of cyanobacteria in perialpine lakes 25 ducted using the same column and eluents of mc/nod analysis. differently from mc target analysis, a 15-min gradient was employed: the starting eluent was 20% b, increased to 90% b at 11 min, and finally restored at 20% b at 15 min. the flow rate was 0.25 ml min−1. the mass detector was operated in the positive electro spray mode (esi+) using the information depended acquisition (ida) mode: a full scan experiment (enhanced mass, ems) in the range 400 1100 da, was used as survey scan; the most intense peaks in ems (with charge state between 1 and 3) were automatically selected and underwent a high-resolution experiment (enhanced resolution, er) and a fragmentation experiment (enhanced product ion). ida threshold was set at 500,000 cps. epi spectra were acquired from 50 to 1000 da with a scan speed of 1000 da s−1 and a collision energy (ce) of 20 v and energy spread (ces) of 10 v. general settings were as follows: ion spray voltage 5000 v, entrance and cell exit potentials 10 v, and interface heater temperature 350°c. the detection of peptidic compounds was enhanced by enabling, in the acquisition software, the built-in algorithm for automatic optimization of ce for each putative peptide according to the specific m/z and charge values. data acquisition and processing were accomplished using analyst 1.5.2 and peakview 2.2 softwares (ab sciex). after acquisition, spectra were manually filtered to get rid of background organic contaminants (by comparison with a blank sample), and only compounds with high quality er and epi data were further considered (a threshold of 30% quality was normally used in peakview). data analysis common and distinctive chemical compounds synthesized by the single species (irrespective of strains) and strains were represented and evaluated by using venn’s diagrams. differences in mass values between species and strains were evaluated by 1-way anova on log-transformed data (sokal and rohlf, 1995). the relationships among the single strains based on the putative assigned compounds found in the untargeted metabolic analysis were evaluated by principal coordinate analysis (pcoa) applied to a dissimilarity matrix computed using the bray & curtis index. the same dissimilarity matrix was used to identify groups of strains by a cluster analysis (ward’s method) (legendre and legendre, 1998). statistical analyses were carried out using the r statistical software (r core team, 2016). results targeted toxin analysis the strains were analyzed with lc-ms/ms methods specifically built and tested for detecting the most common fig. 1. representative chromatogram of a standard mixture of alkaloids. phe is the aminoacid phenylalanine, which is not toxic but is isobaric with atx and can be potentially confused with the toxin. no nco mm er cia l u se on ly l. cerasino et al.26 alkaloid and peptidic toxins with high specificity and sensitivity (cerasino et al., 2016). the results are reported in a schematic view in tab. 2. among the 14 strains, 2 out of 3 strains of t. bourrellyi tested positive for atx: tbour02 and tbour05. the lc-ms chromatogram of tbour05 strain is reported in fig. 2. in the same cultures we could detect two adjacent peaks at about 4.2 min for the transition 412/138 (fig. 2), which could suggest the presence of c3/4 tab. 2. comparative results of the toxin diversity found in the considered strains, obtained with targeted analysis for alkaloids and mcs. in the last column, the number of metabolites found with the untargeted analysis is reported. the presence of c3/4 in strains of t. bourrellyi is suspected but has not been confirmed (see text for details). targeted analysis untargeted analysis species (lake) strain code alkaloids mc (%) metabolites n. aph. flos-aquae (idro) aflos01 19 aflos03 22 aflos04 26 d. lemmermannii (garda) dlemm14 49 dlemm16 74 dlemm21 42 m. aeruginosa (caldonazzo) microc1 lr (96.6), lrdm (3.2), yr (0.2) 36 microc2 lr (90.8), lrdm (9.1), yr (0.1) 46 p. rubescens (garda) prube11 rrdm (83.1), lrdm (16.6), htyrrdm (0.2), rr (0.1) 49 prube17 rrdm (89.4), lrdm (10.1), rr (0.3), htyrrdm(0.1), lr (0.1) 63 prube23 rrdm (99.6), lrdm (0.3), rr (0.1) 39 t. bourrellyi (garda) tbour02 atx, c3/4(?) 27 tbour04 29 tbour05 atx, c3/4(?) 35 fig. 2. chromatogram of the t. bourrellyi strain tbour05, analyzed with the alkaloids-targeted method. the most intense peak (rt 1.0 min) corresponds to atx. the peaks at rt=4.2 min, are approx. 60 times less intense than atx and have been tentatively attributed to the c3/4 toxins. the insert contains the epi (enhanced product ion) spectrum generated by the ion with mass of 412 da. no nco mm er cia l u se on ly metabolic profiles of cyanobacteria in perialpine lakes 27 toxins (tab. 1). however, the lack of the other transitions typical of these toxins (412/332 and 412/314) makes this attribution uncertain. the lack of a specific analytical standard did not allow a certain attribution. looking at data reported in literature about the most common psp (for example dell’aversano et al., 2005), the mass of 412 da could indicate either the [m+h]+ ion of gtx1/4 or the [mso3+h]+ ion of the c3/4 toxins (tab. 1). excluding the gtx1/4, because retention time and fragmentation did not match with those of the pure standard (tab. 1), we are led to hypothesize that the compounds can correspond actually to the c3/4 toxins. the fragmentation pattern of the compounds (reported in the insert of fig. 2), shows two peaks which are consistent with the attribution: 412 and 394 da, respectively attributable to the [m-so3+h]+ and [m-so3h2o+h]+ ions of c3/4. the other intense peaks in the ms/ms spectrum (204, 186 and 138 da) have been described also for other psp (namely stx) and are therefore not diagnostic for a particular molecule, but, instead, can confirm the presence of a similar or identical chemical backbone. the absence of fragments corresponding to the loss of two so3 groups, as in the case of other di-sulfated psp, is however in contrast with this attribution. all the other strains did not show any alkaloids. peptidic toxins were found exclusively in p. rubescens and m. aeruginosa strains (tab. 2). these two species showed a typical mc diversity: rrdm variant was the most abundant in p. rubescens (over 83%), followed by lrdm, htyrrdm, rr and lr. rrdm and lrdm variants together accounted for more than 95.5% of the total mc content. in one strain (prube23), the two less abundant congeners (htyrrdm and lr) were not detected. mc-lr, instead, was the prominent variant in m. aeruginosa strains (more than 90% of the total), with small amounts of two other congeners, namely lrdm and yr (tab. 2). metabolic profiling the untargeted analysis provided a list of metabolites (328 different compounds in total) characterized by a dyad of values: mass and rt (retention time). in the individual strains, we found a variable number of metabolites (tab. 2), from a minimum of 19 to a maximum of 74. most of the peptidic secondary metabolites produced by cyanobacteria were comprised in the considered mass range (400-1100 da). analyzing the distribution of masses, we noted substantial differences among strains (anova, f13,542=13.0, p<0.001; fig. 3) and species fig. 3. boxplot showing the distribution of molecular mass values coming from the untargeted metabolic analysis among the considered 14 strains. boxes represent 25-75th percentiles (first and third quartiles) with median (line in the middle of the rectangle), whereas the top and lower hinges are versions of the first and third quartile computed as in r core team (2016), function boxplot.stats. fig. 4. venn diagram showing the distribution of the compounds found in the untargeted metabolic analysis among the five different cyanobacterial species. the diagram is based on the couples of m/z (mass to charge ratio) and rt (retention time) values. aflos, aphanizomenon flos-aquae; dlemm, dolichospermum lemmermannii; micro, microcystis aeruginosa; prube, planktothrix rubescens; tbour, tychonema bourrellyi. no nco mm er cia l u se on ly l. cerasino et al.28 (groups of strains in fig. 3; anova, f4,551=39.9, p < 0.001). in m. aeruginosa and p. rubescens we found compounds in a wide range of masses, meaning that they produce compounds with very different molecular weights. on the opposite, in aph. flos-aquae and d. lemmermanii, we found that the majority of the compounds were in the lower part of the mass range, meaning they produce mainly low molecular weight compounds. in t. bourrellyi, finally, we found an intermediate situation. when attributing the 328 compounds to the producing species (fig. 4) and to the respective strains (supplementary fig. 1), we found only two compounds common to all species, one having mass 470.2 da (rt 8.12 min), and the other mass 482.3 da (rt 8.80 min). most of the compounds (268 out of 328) were exclusively produced by a single species: d. lemmermannii and p. rubescens had the highest number of exclusive compounds (83 and 70, respectively); m. aeruginosa and t. bourrellyi had lower figures (47 and 44); aph. flos-aquae had the lowest (21). only 62 were the compounds produced by two or more species. as a first attempt to identify some of the compounds, we compared the molecular weights of a subset of the 328 compounds (having molecular masses between 500 and 1200 da; the list is reported in supplementary tab. 1) with those of known compounds. we used online resources (e.g., norine: http://bioinfo.lifl.fr/nrp/) (flissi et al., 2016), and literature (czarnecki et al., 2006; welker et al., 2006; rounge et al., 2007; ersmark et al., 2008; rohrlack et al., 2008; briand et al., 2016; spoof et al., 2016) for finding possible matches. we restricted the search in the mass range typical of cyanobacteria peptidic secondary metabolites: aeruginosins (600-700 da), microginins (700-800 da), anabaenopeptins (800 900 da), microcystins and cyanopeptolins (900-1100 da). based on molecular weight equivalence and consistency of the ms/ms data (fragmentation characteristics and isotopic pattern), we were able to identify 7 potential peptides. for additional 20, an unambiguous attribution was not possible because, although compounds with the same molecular weight were found in databases, either they had not yet been characterized or multiple alternatives were possible; these compounds were generically indicated with the name of the class (aeruginosin, anabaenopeptin, cyanopeptolin, microcystin, etc.) or with “peptide” when the attribution to one class of peptides was not possible (6 compounds) (tab. 3). the molecular weights of these 27 compounds were comprised between approx. 590 and 1180 da. the most represented peptides were cyanopeptolins (7 compounds), followed by anabaenopeptins (6 compounds), and by aeruginosins and microcystins (6 compounds each). interestingly, most of these compounds were found in p. rubescens (17) and m. aeruginosa (10); fewer compounds were found in t. bourrellyi (3) and d. lemmermannii (1), and none in aph. flos-aquae. the pcoa analysis of the distribution of these compounds found in the untargeted metabolic survey allowed us to determine a greater uniformity of metabolites in m. aeruginosa and p. rubescens compared to d. lemmermannii and t. bourrellyi (fig. 5a). however, at a higher level of dissimilarity, the strains tbour05 and dlemm14 were not included in their respective species groups (fig. 5 a,b). fig. 5. a) principal coordinate analysis performed on the distribution of putative secondary metabolites within the cyanobacteria strains. the first and second axes account for the 23% and 21% of the total variance, respectively. the continuous and dashed lines enclose together groups of strains at different level of dissimilarity based on the results of the (b) cluster analysis. no nco mm er cia l u se on ly metabolic profiles of cyanobacteria in perialpine lakes 29 discussion the five cyanobacterial species considered in this investigation are known to produce toxins. in particular, based on the analysis of isolated strains (bernard et al., 2016), aph. flos-aquae and t. bourrellyi were reported as atx producers (sivonen et al., 1989; osswald et al., 2009; salmaso et al., 2015; 2016; shams et al., 2015), whereas m. aeruginosa and p. rubescens were mainly reported as mc producers (metcalf and codd, 2012), and d. lemmermannii as mcs and anatoxin-a(s) producer (sivonen et al., 1992; onodera et al., 1997). our targeted analysis confirmed this behavior for m. aeruginosa and p. rubescens, but not for aph. flos-aquae and d. lemmermanii. this was not unexpected, considered that recent investigations had already shown that the populations isolated in the italian district do not have the capability to produce atx. t. bourrellyi was found to produce atx, as reported also in recent papers (salmaso et al., 2015, 2016; shams et al., 2015). moreover, based on evidences collected in this work, this species was possibly able to synthesize psp toxins, possibly sulfated variants (like c3/4 toxins). m. aeruginosa and p. rubescens strains produced mixtures of different mcs (tab. 2). the most abundant congeners were lr in m. aeruginosa and rrdm in p. rubescens. the relative abundances and identities of congeners were in accordance with previous observations (cerasino and salmaso, 2012; salmaso et al., 2014; cerasino et al., 2016), reporting demethylated mcs (either rr, lr or htyr) as dominant in planktothrix and lr in microcystis. the untargeted analysis revealed the presence in the strains of 328 compounds with molecular mass between tab. 3. list of putative compounds found in the analyzed strains. compounds are ordered according to the molecular mass values. most of the m/z values correspond to single charged [m+h] adducts; in some cases, marked with an asterisk, m/z values correspond to double charged [m+2h] adducts. molecular mass observed m/z and rt putative compound strain diagnostic fragmentation peaks# 592.3 593.3 at 1.37 aeruginosin prube11, prube17 140, 120, 642.4 643.4 at 4.89 aeruginosin 101 dlemm16, tbour05 309, 221, 86 650.4 651.3 at 1.92 aeruginosin 102 prube11, prube23 150, 140, 86, 698.3 699.2 at 0.83 anabaenopeptin tbour04, microc2, prube11 120, 74 714.3 715.3 at 1.23 aeruginosin 126b prube17, prube23 164, 150 816.3 817.3 at 4.82 anabaenopeptin prube11, prube17, prube23 120, 72 830.3 831.3 at 1.09 anabaenopeptin microc1 243, 150, 120 836.4 837.4 at 1.25 anabaenopeptin b prube11 201, 175 850.3 851.3 at 1.57 anabaenopeptin f prube17, prube23 201, 175 855.3 856.3 at 5.03 anabaenopetin microc2 243, 120 983.4 984.4 at 4.06 cyanopeptolin microc1, microc2 243, 150 987.4 988.4 at 7.54 microcystin asp3dhb7-ly prube11 375, 213, 135, 107, 86 996.4 997.4 at 3.21 microcystin (l-meala7)lr prube17 375, 213 997.4 998.4 at 4.61 cyanopeptolin microc1, microc2 243, 150, 120 1008.5 1009.5 at 5.17 microcystin microc1, microc2 375, 213, 135 1010.6 506.8 at 5.41* cyanopeptolin prube17, microc2 243, 215, 150, 120 1011.4 1012.4 at 3.56 cyanopeptolin prube17 243, 150, 120 1023.5 1024.5 at 4.15 cyanopeptolin microc1, microc2 150, 120 1030.5 1031.5 at 5.91 microcystin prube11 375, 213, 135 1039.5 520.7 at 4.20* cyanopeptolin microc2 150 1093.5 1094.5 at 2.53 cyanopeptolin prube11, prube17, prube23, tbour05 150, 107, 84 1107.6 554.8 at 3.60* peptide prube11, prube17, prube23 164, 107, 84 1121.7 561.8 at 4.18* peptide prube23, prube11 339, 164, 107 1123.7 562.3 at 3.04* peptide prube17 150, 120, 84 1163.6 582.8 at 3.85* peptide prube17 164 1179.6 590.7 at 3.40* peptide prube11, prube17 164 1182.7 592.3 at 0.88* peptide microc2 120 #diagnostic fragmentation peaks: 70 pro-immonium, 72 val-immonium, 74 thr-immonium, 84 lys-immonium, 86 leu-immonium, 107 [ch2phoh], 120 phe-immonium, 135 [phch2ch(och3)], 140 choi immonium, 150 metyr, 164 mehty, 175 [arg+h], 201 [arg-co], 213 [glu-mdha+h], 215 [ahp-phe h2o co], 221 [leu-choi], 243 [ahp-phe h2o], 309 [choi-arg nh2+h], 339 [meto-mehty+h], 375 [adda-glu-medha+h] (choi: 2carboxy-6-hydroxyoctahydroindole; mdha:n-methyldehydroalanine; adda: 3-amino-9-methoxy-2,6,8-trimethyl-10-phenyldeca-4,6-dienoic acid); *double charged [m+2h] adducts. no nco mm er cia l u se on ly l. cerasino et al.30 400 and 2000 da. among them, we can have both primary and secondary metabolites. many of these have fragmentation peaks typical of peptides (immonium ions fragments in the lower mass range) in their ms/ms spectra. these compounds could be either ribosomal or non-ribosomal. microcystis and planktothrix exhibit a greater ability to produce molecules having molecular weight above 550 da, as can be clearly seen in fig. 3. this net difference compared to the other species can be related to their remarkable ability to produce different classes of peptidic secondary metabolites (included toxic mcs), as demonstrated by the mass of data available in the literature (czarnecki et al., 2006; welker et al., 2006; rounge et al., 2007; ersmark et al., 2008; rohrlack et al., 2008; briand et al., 2016; spoof et al., 2016). this is confirmed by the fact that, among the 27 peptides reported in tab. 3, the majority have been found in planktothrix and microcystis strains. aphanizomenon and dolichospermum produced much fewer peptides, as their secondary metabolism seems to be more oriented on the production of alkaloids (i.e. atx and hatx). tychonema constitutes an intermediate situation (fig. 3) as it exhibits a higher ability in producing peptides compared to the aphanizomenon and dolichospermum. finally, we need to consider that p. rubescens, t. bourrellyi and aph. flos-aquae have been grown under different light conditions compared to d. lemmermannii and m. aeruginosa; therefore, the metabolism of these two groups of organisms could have been differently influenced by this variable. conclusions the paper allowed the comparison of the toxic potential of five cyanobacterial species common in the lakes of the subalpine italian district (namely aph. flosaquae, d. lemmermannii, m. aeruginosa, p. rubescens, and t. bourrellyi). p. rubescens and m. aeruginosa resulted to be mc producers with typical mc diversities. aph. flos-aquae and d. lemmermannii resulted to be not toxic. finally, t. bourrellyi resulted to produce atx and possibly (at a minor extent) still not fully characterized psp toxins. the comparison of the peptidomic profiles allowed us to classify p. rubescens and m. aeruginosa as extraordinary non-ribosomal peptides producers. a preliminary attempt aimed at identifying the peptidic compounds has revealed that anabaenopeptins and cyanopeptolins were the most represented, but many other peptides are still to be structurally determined. as demonstrated by recent reports (svirčev et al., 2016), the toxic potential of cyanobacteria can be only partially verified by current analytical techniques. efforts to develop more comprehensive but specific methodologies are therefore still needed. acknowledgments we thank the european cooperation in science and technology cost action es1105 cyanocost for networking and knowledge-transfer support. references agha r, lezcano má, labrador mdm, cires s, quesada a, 2014. seasonal dynamics and sedimentation patterns of microcystis oligopeptide chemotypes reveal subpopulations with different ecological traits. limnol. oceanogr. 59:861-871. anas arj, harada k, 2016. evaluation of serine protease inhibitors as potent fviia-stf inhibitors in the blood coagulation cascade. lett. drug des. discov. 13:3-23. bernard c, ballot a, thomazeau s, maloufi s, furey a, mankiewicz-boczek j, pawlik-skowronska b, capelli c, salmaso n, 2016. appendix 2. cyanobacteria associated with the production of cyanotoxins, p. 503-527. in: j. meriluoto, l. spoof and g.a. codd (eds.), handbook on cyanobacterial monitoring and cyanotoxin analysis. wiley & sons, hoboken. blom jf, baumann h i, codd ga, jüttner f, 2006. sensitivity and adaptation of aquatic organisms to oscillapeptin j and [d-asp3, (e)-dhb7]microcystin-rr. arch. hydrobiol. 167: 547-559. briand e, bormans m, gugger m, dorrestein pc, gerwick wh, 2016. changes in secondary metabolic profiles of microcystis aeruginosa strains in response to intraspecific interactions. environ. microbiol. 18:384-400. bubik a, sedmak b, novinec m, lenarčič b, lah tt, 2008. cytotoxic and peptidase inhibitory activities of selected nonhepatotoxic cyclic peptides from cyanobacteria. biol. chem. 389:1339-1346. cerasino l, salmaso n, 2012. diversity and distribution of cyanobacterial toxins in the italian subalpine lacustrine district. oceanol. hydrobiol. stud. 41:54-63. cerasino l, shams s, boscaini a, salmaso n, 2016. multiannual trend of microcystin production in the toxic cyanobacterium planktothrix rubescens in lake garda (italy). chem. ecol. 32:492-506. chorus i, 2012. current approaches to cyanotoxin risk assessment, risk management and regulations in different countries. federal environment agency, dessau-roßlau: 151 pp. codd ga, 1995. cyanobacterial toxins: occurrence, properties and biological significance. water sci. technol. 32:149-156. czarnecki o, henning m, lippert i, welker m, 2006. identification of peptide metabolites of microcystis (cyanobacteria) that inhibit trypsin-like activity in planktonic herbivorous daphnia (cladocera). environ. microbiol. 8:77-87. dell’aversano c, hess p, quilliam ma, 2005. hydrophilic interaction liquid chromatography-mass spectrometry for the analysis of paralytic shellfish poisoning (psp) toxins. j. chrom. a 1081:190-201. ersmark k, del valle jr, hanessian s, 2008. chemistry and biology of the aeruginosin family of serine protease inhibitors. angew. chem. int. ed. 47:1202-1223. fischbach ma., walsh ct, clardy j, 2008. the evolution of no nco mm er cia l u se on ly metabolic profiles of cyanobacteria in perialpine lakes 31 gene collectives: how natural selection drives chemical innovation. proc. natl. acad. sci. usa 105:4601-4608. flissi a, dufresne y, michalik j, tonon l, janot s, noé l, jacques p, leclère v, pupin m, 2016. norine, the knowledgebase dedicated to non-ribosomal peptides, is now open to crowdsourcing. nucl. acids res. 44(d1): d1113-d1118. halme m, rapinoja ml, karjalainen m., vanninen p, 2012. verification and quantification of saxitoxin from algal samples using fast and validated hydrophilic interaction liquid chromatography-tandem mass spectrometry method. j. chrom. b 880:50-57. hamilton dp, salmaso n, paerl hw, 2016. mitigating harmful cyanobacterial blooms: strategies for control of nitrogen and phosphorus loads. aquat. ecol. 50:351-366. iarc, 2010. cyanobacterial peptide toxins, p. 327-412. in: evaluation of carcinogenic risks to humans, vol. 94, ingested nitrate and nitrite, and cyanobacterial peptide toxins, iarc monographs. international agency for research on cancer, lyon. janse i, kardinaal wea, kamst-van agterveld m, meima m, visser pm, zwart g, 2005. contrasting microcystin production and cyanobacterial population dynamics in two planktothrix-dominated freshwater lakes. environ. microbiol. 7:1514-1524. kardinaal wea, janse i, kamst-van agterveld m, meima m, snoek j, mur lr, huisman j, zwart g, visser pm, 2007. microcystis genotype succession in relation to microcystin concentrations in freshwater lakes. aquat. microb. ecol. 48:1-12. komárek j, anagnostidis k, 2005. [cyanoprokaryota. part 2: oscillatoriales]. [süßwasserflora von mitteleuropa, band 19/2.[book in german]. springer spektrum, dodrecht. kotai j, 1972. instructions for preparation of modified nutrient solution z8 for algae. norwegian institute for water research b-11769, oslo: 6 pp. legendre p, legendre l, 1998. numerical ecology. 24. elsevier, amsterdam: 852 pp. metcalf js, codd ga, 2012. cyanotoxins, p. 651-675. in: b.a. whitton (ed.), ecology of cyanobacteria ii. their diversity in time and space. springer, dodrecht. moore r, 1996. cyclic peptides and depsipeptides from cyanobacteria: a review. j. ind. microbiol. 16:134-143. murakami m, suzuki s, itou y, kodani s, ishida k, 2000. new anabaenopeptins, potent carboxypeptidase-a inhibitors from the cyanobacterium aphanizomenon flos-aquae. j. nat. prod. 83:1280-1282. neumann u, forchert a, flury t, weckesser j, 1997. microginin fr1, a linear peptide from a water bloom of microcystis species. fems microbiology lett. 153:475-478. nikolouli k, mossialos d, 2012. bioactive compounds synthesized by non-ribosomal peptide synthetases and type-i polyketide synthases discovered through genome-mining and metagenomics. biotechnol. lett. 34:1393-1403. onodera h, oshima y, henriksen p, yasumoto t, 1997. confirmation of anatoxin�a(s), in the cyanobacterium anabaena lemmermannii, as the cause of bird kills in danish lakes. toxicon 35:1645-1648. osswald j, rellàn s, gago-martinez a, vasconcelos v, 2009. production of anatoxin-a by cyanobacterial strains isolated from portuguese fresh water systems. ecotoxicology 18:1110-1115. paerl hw, huisman j, 2009. climate change: a catalyst for global expansion of harmful cyanobacterial blooms. environ. microbiol. rep. 1:27-37. r core team, 2016. r: a language and environment for statistical computing. r foundation for statistical computing, vienna, austria. url https://www.r-project.org/ rohrlack t, edvardsen b, skulberg r, halstvedt cb, utkilen hc, ptacnik r, skulberg om, 2008. oligopeptide chemotypes of the toxic freshwater cyanobacterium planktothrix can form subpopulations with dissimilar ecological traits. limnol. oceanogr. 53:1279-1293. rounge tb, rohrlack t, tooming-klunderud a, kristensen t, jakobsen ks, 2007. comparison of cyanopeptolin genes in planktothrix, microcystis, and anabaena strains: evidence for independent evolution within each genus. appl. environ. microbiol. 73:7322-7330. salmaso n, capelli c, shams s, cerasino l, 2015. expansion of bloom-forming dolichospermum lemmermannii (nostocales, cyanobacteria) to the deep lakes south of the alps: colonization patterns, driving forces and implications for water use. harmful algae 50:76-87. salmaso n, cerasino l, boscaini a, capelli c, 2016. planktic tychonema (cyanobacteria) in the large lakes south of the alps: phylogenetic assessment and toxigenic potential. fems microbiology ecology. http://www.ncbi.nlm.nih.gov/ pubmed/27402712. salmaso n, copetti d, cerasino l, shams s, capelli c, boscaini a, valsecchi l, pozzoni f, guzzella l, 2014. variability of microcystin cell quota in metapopulations of planktothrix rubescens: causes and implications for water management. toxicon 90:82-96. schindler dw, 2006. recent advances in the understanding and management of eutrophication. limnol. oceanogr. 51:356-363. schopf jw, 2012. the fossil record of cyanobacteria, p. 15-36. in: b.a. whitton (ed.), ecology of cyanobacteria ii: their diversity in space and time. springer, dordrecht. shams s, capelli c, cerasino l, ballot a, dietrich dr, sivonen k, salmaso n, 2015. anatoxin-a producing tychonema (cyanobacteria) in european waterbodies. water res. 69:68-79. shin hj, murakami m, matsuda h, ishida k, yamaguchi k, 1995. oscillapeptin, an elastase and chymotrypsin inhibitor from the cyanobacterium oscillatoria agardhii (nies-204). tetrahedron lett. 36:5235-5238. sivonen k, himberg k, luukkainen r, niemelä si, poon gk, codd ga, 1989. preliminary characterization of neurotoxic cyanobacteria blooms and strains from finland. environ. toxicol. 4:339-352. sivonen k, skulberg om, namikoshi m, evans wr, carmichael ww, rinehart kl, 1992. two methyl ester derivatives of microcystins, cyclic heptapeptide hepatotoxins, isolated from anabaena flos�aquae strain cya 83/1. toxicon 30, 11:1465-1471. sivonen k, borner t, 2008. bioactive compounds produced by cyanobacteria, p. 159-197. in: a. herrero and e. flores (eds.), the cyanobacteria: molecular biology, genomics and evolution. caister academic press. sokal rr, rohlf fj. 1995. biometry: the principles and practices no nco mm er cia l u se on ly l. cerasino et al.32 of statistics in biological research. w. h. freeman: 887 pp. spoof l, blaszczyk a, meriluoto j, ceglowska m, mazurmarzec h, 2016. structures and activity of new anabaenopeptins produced by baltic sea cyanobacteria. mar. drugs 14:8. svirčev z, obradović v, codd ga, marjanović p, spoof l, drobac d, petković a, nenin t, simeunović j, važić t, meriluoto j, 2016. ecotoxicology 25:1353-1363. van apeldoorn me, van egmond hp, speijers gja, bakker gj, 2007. toxins of cyanobacteria. mol. nutr. food res. 51:7-60. welker m, maršálek b, šejnohová l, von doehren h, 2006. detection and identification of oligopeptides in microcystis (cyanobacteria) colonies: toward an understanding of metabolic diversity. peptides 27:2090-2103. welker m, von döhren h, 2006. cyanobacterial peptides nature’s own combinatorial biosynthesis. fems microbiol rev. 30:530-63. yepremian c, gugger mf, briand e, catherine a, berger c, quiblier c, bernard c, 2007. microcystin ecotypes in a perennial planktothrix agardhii bloom. water res. 41:44464456. zurawell rw, chen h, burke jm, prepas ee, 2005. hepatotoxic cyanobacteria: a review of the biological importance of microcystins in freshwater environments. j. toxicol. environ. health b crit. rev. 8:1-37. no nco mm er cia l u se on ly layout 1 introduction current estimates suggest that between 4000 and 5000 phytoplankton species have been described from inland waters (reynolds, 1996, 2006). among them are the prokaryotic phototrophs named cyanoprokaryota (cyanobacteria, cyanophyta or blue-green algae), which are also quantitatively amid the most important organisms on earth (whitton and potts, 2012). their appearance, traced back to the early archaean 3900 ma ago (graham et al., 2009), became a crucial step in the evolution of life in water and, subsequently, on land. the later origin and spread of their heterocytous forms, capable of nitrogen fixation, fits well with the timing of great oxidation event about 2400 ma ago (schopf, 2012). since then cyanoprokaryotes are the only nitrogen fixing organisms that also produce oxygen through photosynthesis, with increasing at steady pace number of known non-heterocytous species that possess this ability (stal, 2012). considering their additional attribute of buoyancy regulation through gas vacuoles, it is to understand why these peculiar organisms for years remained a fascinating topic (reynolds, 2006) in biology, with increasing findings of their great potential as providers of ecosystem services. cyanoprokaryotes are primary colonizers of various (even extreme) habitats, important basis of numerous food chains and have been repeatedly reported for their applications in biotechnology, food industry and pharmacy (whitton, 2012). however, during the last decades the recognition of the group turned in an almost universal contempt (reynolds, 2006) due to their potential toxicity, key role in many harmful blooms and general assuming advances in oceanography and limnology, 2017; 8(1): 131-152 article doi: 10.4081/aiol.2017.6320 this work is licensed under a creative commons attribution-noncommercial 4.0 international license (cc by-nc 4.0). assessment of cyanoprokaryote blooms and of cyanotoxins in bulgaria in a 15-years period (2000-2015) maya p. stoyneva-gärtner,1* jean-pierre descy,2,3 adrien latli,3 blagoy a. uzunov,1 vera t. pavlova,4 zlatka bratanova,4 pavel babica,5,6 blahoslav maršálek,5,6 jussi meriluoto,7 lisa spoof7 1department of botany, faculty of biology, university of sofia “st kliment ohridski”, bld dragan zankov 8, bg-1164, sofia, bulgaria; 2unité d’océanographie chimique, université de liège, sart tilman, b-4000, liège, belgium; 3research unit in organismal biology (urbe), university of namur, 61 rue de bruxelles, 5000 namur, belgium; 4national centre of public health and analyses, str. academik ivan evstratiev geshov 15, 1431 sofia, bulgaria; 5department of experimental phycology and ecotoxicology, institute of botany, czech academy of sciences, lidická 25/27, 602 00 brno, czech republic; 6recetox research centre for toxic compounds in the environment, faculty of science, masaryk university, kamenice 753/5, 625 00 brno, czech republic; 7department of biochemistry, faculty of science and engineering, ǻbo akademi university, turku, finland *corresponding author: mstoyneva@uni-sofia.bg abstract the scientific and public awareness of hazardous photosynthetic prokaryotes (cyanobacteria/cyanoprokaryotes) and especially the contamination of drinking-water reservoirs with cyanotoxins is world-wide increasing. recently much more attention has been paid to the events and results of mass proliferation of these toxic organisms even in south-east european countries in spite of the fact that, as a rule, they are not controlled by national legislation.the present paper presents a summary of results of such studies carried out in summer-autumn periods of the last 15 years (2000-2015) in bulgarian water bodies differing by location, morphometry and trophic status, incl. drinking-water reservoirs, recreational lakes and sites of nature conservation importance. a multivariate analysis allowed to outline the distribution patterns and environmental drivers of the planktonic cyanoprokaryote assemblages in relation with the available data on the water bodies, highlighting species composition and abundance of the main taxa, including potentially toxic species. samples analysis by hplc-dad and/or lc/ms, elisa and in vitro cytotoxicity tests allowed detection of microcystins, nodularins and saxitoxins. toxin concentration ranged between 0.1 and 26.5 µg l–1 in water samples and between 10.9 and 1070 µg g–1 (d.w.) in concentrated (net) samples. despite the fact that microcystins were not found in all studied water bodies and that the recorded levels were still lower in comparison with some other european countries, the fact that cyanotoxins were detected in 16 water bodies (incl. 3 drinking-water reservoirs) could serve as an alert for the need of recognition of cyanotoxins as a new health risk factor in the country. therefore, permanent monitoring with identification of toxins in water bodies at risk and activities for limitation and control of toxic blooms are urgently needed, in combination with increase of the attention to the effects of cyanotoxins on both human health and health of aquatic ecosystems in bulgaria. key words: microcystins; nodularins; saxitoxins; reservoirs; lakes; health risk. received: 30 september 2016. accepted: 19 december 2016. no nco mm er cia l u se on ly m.p. stoyneva-gärtner et al.132 hazardous for human and ecosystem health (carmichael 1994; codd 1994, 1995; chorus and bartram, 1999; codd et al., 1999, 2005a, 2005b; maršálek et al., 2000; huisman et al., 2005; meriluoto and codd, 2005; metcalf and codd, 2012; walker, 2015). the situation in bulgaria is not an exception in this case. moreover, the country is of intermediate position on the balkan peninsula a great european hotspot of biodiversity (griffiths et al., 2004). the studies of its water inhabitants have been conducted since more than hundred years and 216 taxa (incl. 170 cyanoprokaryotes and algae), new for the science, have been described (michev and stoyneva, 2007). however, the inventory assessment of bulgarian water bodies (wbs) brought to the recognition of anthropogenically fast eutrophication with nuisance blooms in rapidly increasing numbers of wbs. this situation was considered by the authors as alarming. since then the number of studies and publications on the cyanotoxins in bulgarian waters, pioneered by pavlova et al. (2006), increased and proved their existence in different sites, some of which of high conservational, drinking-water, sport fishing or recreational importance. a part of the results has been published in bulgarian language, which, in spite of being accepted as one of the official languages of the european union, remains exotic and not understandable for the majority of scientists. therefore, the aim of the present paper is to summarize data on cyanotoxins and water blooms in bulgaria gathered during the last 15 years’ period and to assess them on the background of cyanoprokaryote distribution, diversity and abundance in relation to potential toxin producers and driving environmental variables. methods study area the study area covers practically the whole territory of bulgaria (111,000 km2), a part of the eastern balkan peninsula with the danube and black sea as north and east borders (fig. 1). one-third of the country territory is occupied by plains, while the other is covered by hills, plateaus and higher mountains. its peak mousala (2925 m) is the highest fig.1. map of bulgaria with indication of location of the 120 water bodies studied in 2000-2015. numeration follows tab. 1. no nco mm er cia l u se on ly cyanoprokaryote blooms and cyanotoxins in bulgaria (2000-2015) 133 point of the balkan peninsula, while the lowest parts of bulgaria are at sea level. the climate is temperate continental with a mediterranean influence in its southern (mainly sw) part with a significant amount of snowfall during winter. temperature amplitudes vary in different areas (from -38.3°c to 45.2°c) and precipitation range is from 500 mm in plains to more than 2500 mm in the mountains, being about 630 mm per year on average. due to the interaction of climatic, hydrological, geological and topographical conditions, bulgaria is one of the countries with highest biodiversity in europe (peev et al., 2013). according to the appendix № хі, map а of water framework directive 2000/60/ес (european commission, 2000) bulgaria belongs to the ecoregions № 12 pontic province and № 7 eastern balkans, the borders of which have been updated by cheshmedjiev et al. (2010b). the results from the first inventory of bulgarian wetlands and their biodiversity lead to the conclusion that the country is characterized by great diversity and number of wbs, which cover ca. 1% of its territory (michev and stoyneva, 2007). the term wetlands was consistently used by these authors for all types of wbs, following the definition of ramsar convention on wetlands, signed in 1971. data collection on planktonic cyanoprokaryotes this study compiles and evaluates the currently available data on phytoplankton and relevant environmental data on bulgarian wbs (mainly reservoirs and lakes) gathered during the period 2000-2015 (stoyneva, 2003, 2010, 2014, 2015; beshkova and botev, 2004; kalchev et al., 2004; traykov, 2005; pavlova et al., 2006, 2013a, 2014, 2015; beshkova et al., 2008a, 2008b, 2012; pavlova, 2007; tsanev and belkinova, 2008; teneva et al., 2009, 2010a, 2010b, 2011, 2014; cheshmedjiev et al., 2010a, 2013; stoyanov et al., 2012, 2013, 2016; stoyanov, 2014 stoyneva et al., 2013, 2015; belkinova et al., 2014; dimitrova et al., 2014a, 2014b; dochin and stoyneva, 2014, 2015; dochin, 2015; georgieva et al., 2015). the total number of wbs investigated in the above mentioned studies is 115 (tab. 1, fig. 1). in addition, five recreational lakes have been checked for cyanotoxins, but data on their phytoplankton composition have not been published (pavlova, 2007; pavlova et al., 2013a). they are added in tab. 1 and fig. 1 with numbers 116-120. in tab. 1 we provide the commonly used categories (reservoirs, lakes, swamps, etc.) and the unique number of each wb according to the database of bulgarian wetlands inventory (ibw michev and stoyneva, 2007), where more details on the origin, location, morphometry, trophic status, history, way of exploitation, conservation, etc. could be found. since most investigated wbs are reservoirs, in the text bellow, in an operational way, we assign all wbs to two types – reservoirs (r) and other, mainly natural wbs (owb). according to the geographical location and vertical position all studied wbs could be grouped in the following 8 groups (phyla) of the hierarchical bulgarian wetland classification (michev and stoyneva, 2007): black sea coastal surface lowland wbs (cv.i, 0-200 m asl) – 11 (3 r, 8 owb); black sea coastal surface low mountain wbs (cv.iv; >1000 m asl) – 1 r; inland surface lowland wbs (lv.i; 0-200 m asl) – 41 (35 r, 6 owb); inland surface plain wbs (lv.ii; 200-600(700) m asl) – 24 r; inland surface kettle wbs (lv.iii; 500-1000 m asl– 11 (8 r, 3 owb); inland surface low mountain wbs (lv.iv; >1000 m asl) – 10 r; inland surface middle mountain wbs (lv.v; 1000-1800 m asl – 3 (2 r, 1 owb); inland surface high mountain and alpine wbs (lv.vi; >1800 m asl – 19 (2 r, 17 owb). among the wbs studied 36 are of conservational importance (tab. 1): 21 are included in the red list of bulgarian wetlands as critically endangered (cr) – 9, endangered (en) – 6 and vulnerable (vu) – 6 (michev and stoyneva, 2005, 2007); 36 belong to protected areas with different international and national status (michev and stoyneva, 2007); 17 are special subjects of national action plan for conservation of wetlands of high significance of bulgaria for 2013-2022 (vassilev et al., 2013). phytoplankton sampling and laboratory processing phytoplankton sampling procedures and measurements of the main environmental variables (water temperature t, ph, electric conductivity cond, secchi depth sd, total phosphorus tp and total nitrogen -tn) were based on international and bulgarian standards (e.g., cen en 15204, 2006; state order n 4/14.09.2012, 2013; belkinova and gecheva, 2013 and references therein). some studies included more environmental data (chlorophyll a, dissolved oxygen, saturation, no3, po4, etc.) with tools and protocols described in detail in the relevant publications, but all of them generally followed the same design. more differences concern the microscopic processing of the samples because the cited standards provide possibilities to choose between inverted microscopy and standard microscopy in combination with different types of counting chambers. in the works of the authors of this paper standard light microscopy was used in combination with thoma or burker counting chambers. always cell was the main counting unit. the biomass estimations were based on the measurement of the dimensions of each cell according to the method of stereometrical approximations (rott, 1981) instead of the often recommended and broadly applied by other bulgarian authors usage of average cell sizes. the reasons for choosing this more time-consuming way of work were described in stoyneva et al. (2015) but for the purpose of the present paper they could be briefly summarised as follows: i) necessity of cell measurements for correct taxa identification; ii) differences in cell size during the cell division process; 3) variations in cell size of the same taxon in different wbs, no nco mm er cia l u se on ly m.p. stoyneva-gärtner et al.134 tab. 1. bulgarian wbs studied in the period 2000-2015. the names are provided according to the bulgarian wetlands inventory (ibw) through transliteration of their vernacular names; in brackets the synonyms used in the literature. number name type gl az pa rl nap ibwnumber 1 aheloy* r cv.iv. 3 ibw3032 2 aleksandrovo r lv.i. 3 ibw2017 3 aleksandur stamboliyski* r lv.i. 3 ibw2056 4 blato alepu* s cv.i. 1 x cr ibw1770 5 musalensko ezero 3 (alekovo ezero) l lv.vi. 13 x vu ibw0078 6 antimovo* r lv.i. 3 ibw 2818 7 asenovets r lvii 4 ibw2549 8 asparuhov val* r lv.i. 2 ibw 3674 9 atanasovsko ezero* l cv.i. -1 x vu x ibw1900 10 baniska r lv.i. 3 ibw9042 11 batak* r lv.v. 8 ibw1316 12 bebresh* r lv.ii. 5 ibw2397 13 belmeken r lv.vi. 10 ibw1187 14 blatse do aec-belene ts lv.i. 1 x ibw4638 15 beli iskur* r lv.vi. 10 x ibw1180 16 beli lom r lv.ii. 5 ibw2810 17 bezbozhsko ezero 1 (ezero bezbog)* l lv.vi. 11 x ibw0442 18 balastrierni ezera bistratsite/bistraka l lv.iii. 5 ibw4563 19 bistritsa* r lv.iii. 6 ibw1067 20 blatse do aec-kozloduy t lv.i. 1 21 borovitsa r lv.iv. 7 ibw1580 22 boyka r lv.ii 4 ibw2573 23 vaya (burgasko ezero)* l cv.i. 1 x cr x ibw0191 24 barzina* r lv.i. 3 ibw1276 25 nevenino ezero 1 (chernoto ezero)* l lv.vi. 12 x ibw0371 26 chirpan* r lv.i. 3 ibw1704 27 choklyovo blato* s lv.v. 7 x vu x ibw 0003 28 gergiysko ezero 2* l lv.vi. 12 x ibw0480 29 gorni dabnik* r lv.i. 3 ibw 5606 30 daskal atanasovo r lv.i. 3 ibw2219 31 devets (monchovets) r lv.ii 4 ibw10869 32 dospat* r lv.v. 5 ibw3155 33 drenovets r lv.i. 3 ibw1128 34 dabnika r lv.ii. 5 ibw5393 35 durankulashko ezero (durankulak)* l cv.i. 1 x cr x ibw0216 36 dyakovo* r lv.iv. 6 ibw1033 37 blato dyuleva bara t lv.i. 1 x en x ibw0154 38 eleshnitsa* r lv.i. 2 ibw3023 39 enitsa r lv.i. 3 ibw1444 40 ivaylovgrad* r lv.i. 3 ibw2271 41 iskur* r lv.iii. 7 ibw1200 42 hr. smirnenski na reka lom* r lv.i. 3 ibw1135 43 hr. smirnenski na reka yantra (hr. smirneski/gabrovo)* r lv.ii. 5 ibw2080 44 kamenets r lv.i. 3 ibw2162 45 karaisen r lv.i. 3 ibw5113 46 musalensko ezero 7 (karakashevo ezero) l lv.vi. 13 x vu ibw0080 47 seyatchi (kavacite, popovo)* r lv.ii. 4 ibw2606 48 kamchiya* r lv.iv. 5 ibw2745 49 koprinka* r lv.ii. 5 ibw2062 50 kovachitsa r lv.i. 3 ibw1160 51 krapets* r lv.iii. 5 ibw2000 52 kremensko ezero 2* l lv.vi. 12 x ibw9088 to be continued on next page no nco mm er cia l u se on ly cyanoprokaryote blooms and cyanotoxins in bulgaria (2000-2015) 135 tab. 1. continued from previous page. number name type gl az pa rl nap ibwnumber 53 krichim r lv.i. 3 ibw1366 54 krushovitsa 3 (krushovitsa) r lv.i. 1 ibw1452 55 kula r lv.ii. 4 ibw1105 56 kardzhali* r lv.ii. 5 ibw1668 57 musalensko ezero 1 (ledeno ezero) l lv.vi. 13 x vu ibw0076 58 lazhenska bara (ladzhenska bara) r lv.i/ 2 ibw2166 59 mandra* r cv.i. 1 x en x ibw1720 60 marichino ezero 2 (gorno marichino ezero) l lv.vi. 12 x ibw0085 61 marichino ezero 3 (dolno marichino ezero) l lv.vi. 12 x ibw0086 62 murtvo blato ts lv.i. 1 x en x ibw0158 63 ezero momin brod* l lv.i. 1 ibw8307 64 novo zhelezare r lv.ii. 4 ibw1475 65 ognyanovo* r lv.iii. 6 ibw2340 66 ogosta* r lv.i. 3 ibw 1137 67 ovcharitsa r lv.i. 3 x x ibw2317 68 ovchi kladenets r lv.i. 3 ibw2367 69 pchelina* r lv.iii. 6 ibw1039 70 blato peschina (pischina) t lv.i. 1 x en x ibw0156 71 poletkovtsi 2 (poletkovtsi)* r lv.ii. 4 ibw1103 72 pomoriysko ezero* l cv.i. -1 x vu x ibw 0189 73 popovo ezero 2 (popovo ezero)* l lv.vi. 11 x ibw0447 74 poroy* r cv.i. 2 ibw3038 75 pyasuchnik 1 (pyasuchnik)* r lv.ii. 4 x ibw5834 76 rabisha* r lv.ii. 4 ibw1102 77 rasovo 2 (rasovo) r lv.i. 3 ibw1158 78 redzhepsko ezero 2 (redzhepsko ezero) r lv.vi. 12 x ibw0342 79 dragash voyvoda r lv.i. 1 x ibw3935 80 onogour (efreytor bakalovo) r lv.i. 3 x ibw5667 81 bunderishko ezero 9 (ribno ezero)* l lv.vi. 11 x ibw0400 82 shablensko ezero (shabla)* l cv.i. 1 x cr x ibw0219 83 shablenska tuzla l cv.i. -1 en ibw0218 84 kayabash 2 (golyamo skalensko ezero) r lv.iv. 5 ibw2659 85 kayabash 1 (malko skalensko ezero) r lv.iv. 5 ibw2658 86 hisar 12 (sinyata reka)* r lv.ii. 5 ibw1893 87 sopot* r lv.ii. 5 ibw1437 88 blato srebarna* l lv.i. 1 x en x ibw0208 89 srechenska bara* r lv.iv. 5 ibw3668 90 stoychovtsi (stoykovtsi) r lv.iv. 8 ibw3237 91 studena* r lv.iv. 7 ibw 1060 92 studen kladenets* r lv.ii. 4 ibw1763 93 suedinenie* r lv.i. 3 ibw2642 94 telish* r lv.i. 3 ibw1413 95 ticha* r lv.i. 3 ibw2700 96 toshkov chark r lv.v. 9 x ibw1315 97 trakiets r lv.ii. 4 ibw1677 98 tri kladentsi* r lv.i. 3 ibw1275 99 tsonevo* r lv.i. 2 ibw3022 100 varnensko ezero* l cv.i. 1 x cr ibw0203 101 vlahinsko ezero 1* l lv.vi. 12 x ibw0475 102 vucha r lv.i. 3 ibw 3143 103 valchovets r lv.i. 2 ibw2129 104 yarlovets (yarlovtsi) r lv.iv. 7 ibw 1038 105 yasna polyana* r cv.i. 2 ibw2887 to be continued on next page no nco mm er cia l u se on ly m.p. stoyneva-gärtner et al.136 or even in the same water body due to different temperature, nutrient content or grazing pressure (stoyneva et al., 2007). we believe that the results obtained in this way reflect the real biomass for a given time and site, especially when the samples are analysed by the same person. multivariate analyses considering the reasons stated above, after analyzing of all available, but quite heterogenous data on the 20002015 phytoplankton in bulgarian wbs (references provided in the “data collection” paragraph above), we chose to build a homogenous data set (stoyneva, 2014, 2015; this study) by including only the wbs (sites) with at last 5 samples processed per site, for which the environmental data listed above were available. wbs in which cyanotoxins have been investigated by the authors of this paper, have been included in the dataset. in addition, wbs were chosen to represent all 8 general groups (from cv.i to lv.vi) proposed in the bulgarian wetlands classification and the cr, en and vu conservational threat categories (michev and stoyneva 2005, 2007) as well. in this way, the final dataset comprised 61 wbs (marked with asterisk* in tab. 1) and average values of biomass of 93 species, varieties and forms grouped in 24 genera, environmental data, total cyanoprokaryote biomass (tbc), total phytoplankton biomass (tbs). the dataset also includes average data biomass of other phytoplankton dominant groups, as well as data on wbs morphometry (area and depth), geographic location in the 8 groups of michev and stoyneva (2007), exact altitude asl (m) or as 14 elevation classes proposed in the ibw database of michev and stoyneva (2007): -1 – below 0; 1-0-50; 2 – 50-100; 3 – 100-200; 4 – 200-300; 5 – 300-500; 6 – 500-700; 7 – 700-1000; 8 – 1000-1100; 9 – 1100-1600; 10 – 16002000; 11 – 2000-2300; 12 – 2300-2700; 13 – >2700 m asl. multivariate analyses were run using principal components analyis (pca) and redundancy analysis (rda), using the r-software (r 3.1.2 version, r development core team, 2010) and the “ade4” package (thioulouse et al., 1997). the aims were i) to identify the main environmental gradients among the samples of lakes/reservoirs and ii) to investigate the response of the cyanoprokaryote assemblages (at the genus level) to these gradients. all variables were normalised; the cyanobacterial abundances were transformed using the hellinger transformation (legendre and gallagher, 2001). taxonomic sources, terminology and biodiversity assessment taxonomic sources include mainly the standardly used volumes of middle european freshwater flora with some published updates (komárek 2013; komárek and anagnostidis 1999, 2005; komárek et al. 2011, etc.) and algaebase (guiry and guiry 2016). the latin names follow the above mentioned botanical sources and the international code of the nomenclature for algae, fungi and plants (mcneil et al., 2012). therefore, the term cyanoprokaryotes is used in tab. 1. continued from previous page. number name type gl az pa rl nap ibwnumber 106 yastrebino* r lv.ii. 5 ibw2602 107 shilkovtsi (yovkovtsi)* r lv.ii. 5 ibw2105 108 zhrebchevo* r lv.ii. 4 ibw2545 109 lomtsi r lv.ii. 4 ibw2772 110 pancharevo r lv.iii. 6 ibw1088 111 zhernov r lv.i. 2 ibw4639 112 ezero bliznaka l lv.vi. 11 x cr x ibw0350 113 ezero bubreka l lv.vi. 11 x cr x ibw0349 114 ezero okoto l lv.vi. 12 x cr x ibw0348 115 ezero sulzata l lv.vi. 12 x cr x ibw0347 116 balastrierni ezera dolni bogrov l lv.iii. 6 ibw0708 117 botunets r lv.iii. 6 ibw1792 118 ezero 1 v kvartal druzhba l lv.iii. 6 ibw5284 119 krasava r lv.iv. 7 ibw1049 120 rudnichno ezero кutina 1 l lv.iii. 6 ibw0705 *denote wb included in the dataset for the principal component analysis (pca) and redundancy analysis (rda) for this paper. r, reservoir; l, lake; s, swamp; t, temporary swamp. for each wb, the main group of geographic location and vertical position (gl) is shown (cv.i – lv.vi) and altitude zone (az) in 14 elevation classes (-1 -13); for details on gl and az see the text. the conservation value of the wb is provided as protected area (pa), national action plan for conservation of wetlands of high significance of bulgaria for 2013-2022 (nap) and red list of bulgarian wetlands (rl); cr, critically endangered; en, endangered; vu, vulnerable. ibw number is the number of wb in the bulgarian wetlands inventory (michev and stoyneva, 2007), where more details on them could be found. no nco mm er cia l u se on ly cyanoprokaryote blooms and cyanotoxins in bulgaria (2000-2015) 137 the paper instead of cyanobacteria. for each species and genus, the frequency quotient (fq) of occurrence in all studied sites was estimated based on its presence / absence in each site. the frequency quotients were grouped in classes with a step of 10% (i class – 0-10%, ii class – 1020%, etc.). the evaluation of cyanoprokaryote diversity was done as comparisons with the recently estimated data on the total algal biodiversity in bulgaria (stoyneva, 2014) and the total diversity of cyanoprokaryota in the country (stoyneva et al., 2016). collection and assessment of cyanotoxin data, algal blooms and potential toxin producers the assessment of cyanotoxins registered in the country in relation to their potential producers is based on the works of pavlova (2007), pavlova et al. (2006, 2007, 2013a, 2014, 2015), teneva et al. (2009, 2010a, 2010b, 2011, 2014), stoyanov et al. (2012) and georgieva et al. (2015) – tabs. 2 and 3. for this study the biomass of the species published earlier by pavlova et al. (2006, 2014) with their cell numbers per liter (or milliliter) was estimated (tabs. 2 and 3). results on toxin findings were superimposed on the data on phytoplankton blooms according to the references in tabs. 2 and 3. the cyanotoxin concentrations were evaluated according to the who standards (1998, 2003) since their maximum acceptable levels are not indicated in bulgarian national legislation (pavlova et al., 2013b). results diversity and abundance of cyanoprokaryotes a total of 210 taxa (207 species, 1 variety and 2 forms) from 69 genera of cyanoprokaryota were recorded in the 2000-2015 period. the distribution of species in the three orders chroococcales, oscillatoriales and nostocales clearly shows the chroococcales (85) as the richest order and the better representation of non-heterocytous filamentous forms (69) in comparison with hetereocytous, filamentous taxa (56). at the genus level, heterocytous taxa (17) were also less numerous than non-heterocytous filamentous (21) and than coccal taxa (31). the most speciesrich genera were dolichospermum (11), microcystis (11), anabaena (9), aphanocapsa (9), oscillatoria (9), chroococcus (8), phormidium (8), pseudanabaena (8), romeria (8), anabaenopsis (6) and aphanizomenon (6). most of the species were rare (96 were found only in one site) and fq ranged from 1 to 34%. among the 19 broadly distributed algae (in ≥11 sites), 14 have been identified at species level (fig. 2), with aphanizomenon flosaquae ralfs ex bornet & flahault being the most widespread taxon. most of the 69 genera were rare (18 were found in one site), with fq range from 1 to 43%. the 22 genera with the widest distribution (in ≥12 sites) are shown on fig. 3: aphanizomenon is outstanding as the most widespread genus in fig. 2. distribution of the most widespread cyanoprokaryotes in the phytoplankton of bulgarian water bodies in the period 2000-2015. 0-40, number of water bodies in which species were found; *species found in the samples containing cyanotoxins. no nco mm er cia l u se on ly m.p. stoyneva-gärtner et al.138 ta b. 2 .c ya no to xi ns in b ul ga ri an w b s, o rg an iz ed a cc or di ng to th e fi rs t y ea r o f f in di ng . w at er b od y m ai n us e d at e b io m as s m c (t yp es ) m c to ta l n o d a n ta s t x m et ho ds s ou rc e l r r r y r l a y r eq m c /n o d n o d h p l c e lis a i v c t b is tr its a r s 2 0. 8. 20 04 1 .5 3* 0 .8 6 x x 1. 64 x pa vl ov a et a l. 20 06 14 .1 0. 20 13 n. a. 0 .6 0 .6 x pa vl ov a et a l. 20 15 b or ov its a d r 2 6. 7. 20 06 n ot in d. 0 .0 90. 12 x x x t en ev a et a l. 20 10 b 2 6. 9. 20 06 n ot in d. x x x x 0 .1 8 2 .5 x x t en ev a et a l. 20 10 b v uc ha w s a ug 20 08 0 .1 1 0 .0 04 x x x te ne va e t a l. 20 10 a s ep t2 00 9 0 .0 62 0 .0 05 x x x te ne va e t a l. 20 10 a st ud en k la de ne ts -d am w al l p ar t ir ju ly 20 08 n ot in d. 0 .2 -0 .4 x x te ne va e t a l. 20 11 a ug 20 08 0 .0 7 0 .2 -0 .4 x x t en ev a et a l. 20 11 s ep t2 00 8 0 .1 0 .2 -0 .4 x x t en ev a et a l. 20 11 j ul y2 00 9 0 .0 21 0 .2 -0 .4 x t en ev a et a l. 20 12 a ug 20 09 0. 01 1 0 .2 -0 .4 x x x t en ev a et a l. 20 11 s ep t2 00 9 0 .0 33 0 .2 -0 .4 x <0 .1 0 x x te ne va e t a l. 20 11 st ud en k la de ne ts ta il pa rt j ul y2 00 8 n ot in d. 0 .2 -0 .4 x x te ne va e t a l. 20 11 a ug 20 08 0 .1 9 0 .2 -0 .4 x x t en ev a et a l. 20 11 s ep t2 00 8 0 .1 0 .2 -0 .4 x x t en ev a et a l. 20 11 j ul y2 00 9 0 .0 94 x < 0. 10 x x t en ev a et a l. 20 11 a ug 20 09 0. 76 1 x < 0. 10 x x t en ev a et a l. 20 11 s ep t2 00 9 0 .4 53 0 .2 -0 .4 x x x te ne va e t a l. 20 11 tr ak ie ts d r ju ly 20 08 0 .5 x x x x 0. 09 x x x x te ne va e t a l. 20 09 a ug 20 08 0 .0 4 x x x x 0 .0 18 x 0 .0 1 x x x te ne va e t a l. 20 09 s ep t2 00 9 0. 06 x x x x 0 .0 14 x x x x te ne va e t a l. 20 09 k ru sh ov its a r s 4 .1 0. 20 09 2. 32 1 x x x s to ya no v et a l. 20 12 e ni ts a r s 4 .1 0. 20 09 1 4. 57 2 x x x s to ya no v et a l. 20 12 v al ch ov et s w s 4 .1 0. 20 09 0 0 .0 2 x x x s to ya no v et a l. 20 12 pc he lin a r s 3. 8. 20 11 0. 16 7 0 .5 0 .5 x pa vl ov a et a l. 20 14 2 .1 0. 20 14 6 .2 0 .1 7 0 .0 7 0. 13 0. 37 x pa vl ov a et a l. 20 15 k ay ab as h 2 ir ju ly 20 11 3 0 0 .4 x x x te ne va e t a l. 20 14 s ep t2 01 1 6 5. 42 0. 4 x x x te ne va e t a l. 20 14 st ud en a d r 2 8. 9. 20 11 0. 1 0 .1 x pa vl ov a et a l. 20 14 14 .1 0. 20 13 0. 1 0 .1 p av lo va e t a l. 20 15 5 .1 1. 20 15 0 .0 04 * 0 .2 0 .4 0 .6 x g eo rg ie va e t a l. 20 15 e ze ro m om in b ro d r s 1 9. 8. 20 12 0 .8 1 0 .3 1 .3 x p av lo va e t a l. 20 15 21 .1 0. 20 13 3 .7 1 1 0 .4 0 .4 1 .8 x pa vl ov a et a l. 20 15 d ur an ku la sh ko e ze ro r s 1 8. 8. 20 13 10 .5 0 .2 0 .1 0 .9 1 .2 x pa vl ov a et a l. 20 15 2 9. 7. 20 11 7. 01 8 .3 12 .7 5 .5 2 6. 5 x p av lo va e t a l. 20 14 1 8. 8. 20 13 10 .5 0 .2 0 .1 0 .9 1 .2 x pa vl ov a et a l. 20 15 n et s am pl es / sc um “ bi om as se s” v ay a r s 5. 8. 20 04 6 8. 56 * 1 4 x x 4 2 x pa vl ov a et a l. 20 06 v ay a -q ua y 5 .8 .2 00 4 3 2. 34 * 2 60 x x 1 07 0 x pa vl ov a et a l. 20 06 d ur an ku la sh ko e ze ro r s 3. 8. 20 04 5. 43 * 2 60 11 0 x pa vl ov a 20 07 ; p av lo va et a l. 20 06 . 2 00 7; 1 4. 7. 20 05 n .a . 2 04 x x 51 7 x p av lo va e t a l. 20 07 , 2 01 3 2 9. 7. 20 11 n .a . 2 7. 5 2 2. 1 x( tr ac e) 49 .6 p av lo va e t a l. 20 14 1 8. 8. 20 13 n .a . 6 3. 5 10 3. 2 4 7. 6 2 14 .3 x pa vl ov a et a l. 20 15 to b e co nt in ue d on n ex t p ag e no nco mm er cia l u se on ly cyanoprokaryote blooms and cyanotoxins in bulgaria (2000-2015) 139 the country during the analyzed period. in this study, we classified as picoplankton single spherical cells of dimensions 0.5-1.5 µm, commonly referred as pcy (picocyanobacteria) in contrast to the second morphological group of colonial picoplankters – cpcy (colonial picocyanobacteria) – stockner et al. (2002), callieri et al. (2012). the number of species and genera per site ranged between (0)1 and 85, and 1 to 35, respectively (fig. 4), the highest number detected in the shallow lakes blato srebarna (site 88: 85/35) and vaya (site 23: 63/30). the comparison of distribution of species and genera per site, expressed on fig. 4, shows that in most of the wbs each genus is represented by a single species, which makes the discussion of the distribution on generic level and further pca analysis quite reasonable. the contribution of the cyanoprokaryotes to the total phytoplankton biomass in different wbs and in different sampling periods ranged from 0 to 100%, with values exceeding 65-75 mg l–1 in summer periods (e.g., in the coastal lake vaya stoyneva, 2003; pavlova et al., 2006, 2007; dimitrova et al., 2014a) and reaching 95.9 mg l–1 in average for the period 2010-2014 (this study). the assessment of the average cyanoprokaryote biomass in 61 wbs from the dataset shows that in four sites (the lakes vaya and blato srebarna, and the reservoirs mandra and tri kladentsi), its values were over who’s (2003) threshold for medium health risk category (10 mg l–1) in recreational waters and in five other (the lakes durankulashko ezero and ezero momin brod, and the reservoirs asparuhov val, chirpan and pchelina) from the rest 57 wbs, they were over the low health risk category threshold (2 mg l–1), as they were re-estimated from cell numbers in biomass values by mishke et al. (2011). the single investigation of steady-states, as sommer and padisák (1993) have defined them, outlined their presence in the period analyzed for this paper only in the lake vaya: 3 weeks dominance (89%) of microcystis wesenbergii (komárek) komárek in kondratieva, aphanizomenon flos-aquae and dolichospermum spiroides (klebahn) wacklin et al. in august–september 2001 and 4 weeks dominance (98%) of m. wesenbergii and aphanizomenon gracile (lemmermann) lemmermann in augustseptember 2002 (stoyneva, 2003). the lake vaya is the only wb for which the total carbon content was estimated (dimitrova et al., 2014a). its mean value of 9.7 mg l–1 (2004-2006) together with the average biomass of 46 mg l–1 for the same period confirmed the hypertrophic status of the lake. cyanoprokaryotes dominated constantly in the total carbon content, reaching absolute maxima of 25.1 and 26.9 mg l–1 in august 2005 and 2006. dominance of this group with water blooms (up to 200 mg l–1 ) was detected also in more recent studies of this shallowest coastal lake and, in parallel, cyanoprokaryote blooms were documented for 19 more bulgarian wbs: alepu, durankulashko ezero,ta b. 2 .c on tin ue d fr om p re vi ou s pa ge . w at er b od y m ai n us e d at e b io m as s m c (t yp es ) m c to ta l n o d a n ta s t x m et ho ds s ou rc e l r r r y r l a y r eq m c /n o d n o d h p l c e lis a i v c t n et s am pl es / sc um “ bi om as se s” m an dr a ir . r s 6. 8. 20 04 2. 61 * 35 x x 6 3 x pa vl ov a et a l. 20 06 pc he lin a ir 19 .8 .2 00 4 0 .0 8* 1 40 x x 53 6 x pa vl ov a et a l. 20 06 3 .8 .2 01 1 n .a . 1 1. 14 11 .1 4 x p av lo va e t a l. 20 14 3 1. 7. 20 12 n .a . x (t ra ce s) x p av lo va e t a l. 20 15 1 8. 9. 20 12 n .a . x (t ra ce s) x p av lo va e t a l. 20 15 2 .1 0. 20 14 n .a . 1 32 .8 9 1. 5 71 .8 2 96 .1 x p av lo va e t a l. 20 15 sh ab le ns ko e ze ro r s 3. 8. 20 04 0. 62 * 40 4 0 x p av lo va e t a l. 20 06 . 2 00 7 1 4. 7. 20 05 n .a . 2 83 x x x 1 01 8 x p av lo va 2 00 7; p av lo va e t a l. 20 07 sh ab le ns ko e ze ro qu ay 3 .8 .2 00 4 1 1. 97 * 1 4 1 4 x p av lo va e t a l. 20 06 . 2 00 7 1 4. 7. 20 05 n .a . 2 92 x x x 98 2 x p av lo va 2 00 7; p av lo va e t a l. 20 07 e ze ro m om in b ro d r s 1 9. 8. 20 12 n .a . 23 4 2 7 x pa vl ov a et a l. 20 15 st ud en a d r 1 4. 10 .2 01 3 n .a . 8. 1 2 0 .8 10 .9 x pa vl ov a et a l. 20 15 *b io m as s es tim at ed fo r t hi s st ud y; x , p re se nc e / n ot v al ue in di ca te d; iv c t, in v itr o cy to to xi ci ty ; n .a ., no t a na ly se d; n ot . i nd ., no in fo rm at io n in th e pu bl ic at io n; b io m as s [m g l –1 ], re la te d to th e cy an op ro ka ry ot es in th e re le va nt s am pl es ; d r , d ri nk in gw at er r es er vo ir ; w r , r es er vo ir f or w at er s up pl ; i r , r es er vo ir f or ir ri ga tio n an d/ or e ne rg y su pp ly ; r s, r ec re at io na l s ite . c on ce nt ra tio ns a re e xp re ss ed a s µg l –1 (w at er bo dy ) o r i n µg g –1 dr y w ei gh t ( bi om as s) . no nco mm er cia l u se on ly m.p. stoyneva-gärtner et al.140 tab. 3. water samples with negative results for cyanotoxins in bulgarian wbs (2000-2015). water body -water samples checked for toxins with date methods source negative results/organized by years of the investigations hplc elisa in vitro ct botunets 20.8.2004 x pavlova et al. 2006 ezeretsko ezero 3.8.2004 x pavlova et al. 2006 2005 x pavlova 2007; pavlova et al. 2013a choklyovo blato 19.8.2004 x pavlova et al. 2006 2005 x pavlova 2007; pavlova et al. 2013a balastrierni ezera dolni bogrov 20.8.2004 x pavlova et al. 2006 ezero 1 v kvartal druzhba 27.8.2004 x pavlova et al. 2006 iskur 18.8.2004 x pavlova et al. 2006 2005 x pavlova 2007; pavlova et al. 2013a rudnichno ezero kutina 1 20.8.2004 x pavlova et al. 2006 studena 19.8.2004 x pavlova et al. 2006 2005 x pavlova 2007; pavlova et al. 2013a 31.7.2012 x pavlova et al. 2015 3.8.2011 x pavlova et al. 2014 18.9.2012 x pavlova et al. 2015 12.8.2013 x this study 2.10.2014 x pavlova et al. 2015 4.8.2015 x georgieva et al. 2015 yasna polyana 6.8.2004 x pavlova et al. 2006 2005 x pavlova 2007; pavlova et al. 2013a bistritsa 2005 x pavlova 2007; pavlova et al. 2013a 3.8.2011 x pavlova et al. 2014 28.9.2011 x pavlova et al. 2014 31.7.2012 x pavlova et al. 2015 18.9.2012 x pavlova et al. 2015 12.8.2013 x pavlova et al. 2015 2.10.2014 x pavlova et al. 2015 4.8.2015 georgieva et al. 2015 5.11.2015 x georgieva et al. 2015 borovitsa 2005 x pavlova 2007; pavlova et al. 2013a vaya 2005 x pavlova 2007; pavlova et al. 2013a 3.8.2011 x pavlova et al. 2014 2.8.2012 x pavlova et al. 2015 16.8.2012 x pavlova et al. 2015 21.8.2013 x pavlova et al. 2015 durankulashko ezero 2005 x pavlova 2007; pavlova et al. 2013a 12.7.2012 x pavlova et al. 2015 8.8.2012 x pavlova et al. 2015 krasava 2005 x pavlova 2007; pavlova et al. 2013a hristo smirnenski na reka yantra 2005 x pavlova 2007; pavlova et al. 2013a yovkovtsi (vt) 2005 x pavlova 2007; pavlova et al. 2013a mandra 2005 x pavlova 2007; pavlova et al. 2013a pchelina 2005 x pavlova 2007; pavlova et al. 2013a 28.9.2011 x pavlova et al. 2014 31.7.2012 x pavlova et al. 2015 18.9.2012 x pavlova et al. 2015 12.8.2013 x pavlova et al. 2015 14.10.2013 x pavlova et al. 2015 4.8.2015 georgieva et al. 2015 5.11.2015 x georgieva et al. 2015 ticha 2005 x pavlova 2007; pavlova et al. 2013a trakiets 2005 x pavlova 2007; pavlova et al. 2013a shablensko ezero 2005 x pavlova 2007; pavlova et al. 2013a kayabsh 1 july2011 x x x teneva et al. 2014 sept2011 x x x teneva et al. 2014 july2012 x x x teneva et al. 2014 sept2012 x x x teneva et al. 2014 kayabash 2 july2012 x x x teneva et al. 2014 sept2012 x x x teneva et al. 2014 ezero momin brod 24.5.2013 x this study no nco mm er cia l u se on ly cyanoprokaryote blooms and cyanotoxins in bulgaria (2000-2015) 141 blato srebarna and for the reservoirs acheloy, boyka, borovitsa, barzina, daskal atanasovo, enitsa, kamenets, krushovitsa, kardzhali, mandra, ovchi kladenets, pchelina, seyachi, suedinenie, tri kladentsi, vucha (traykov, 2005; cheshmedjiev et al., 2010a; stoyanov et al., 2012; belkinova et al., 2014; stoyneva, 2014) (fig. 1). the same authors and also dochin and stoyneva (2014, 2015) indicated the “presence of toxic species (anabaena, aphanizomenon, microcystis, etc.)” for the following 43 wbs: choklyovo blato, durankulashko ezero, shablensko ezero, blato srebarna and for the reservoirs aheloy, aleksandrovo, antimovo, asparuhov val, batak, boyka, barzina, daskal atanasovo, dospat, drenovets, dabnika, dyakovo, hristo smirnenski (na reka lom), ivaylovgrad, kamenets, krapets, krushovitsa, koprinka, kovachitsa, kula, lomtsi, mandra, ogosta, ovcharitsa, pancharevo, pchelina, poletkovtsi, poroy, pyasuchnik, rabisha, rasovo, seyachi, suedinenie, telish, tri kladentsi, valchovets, vucha, yastrebino and zhrebchevo (fig. 1). toxic species were indicated as blooming in the reservoirs kamenets, beli lom and ovchi kladenets but have not been enlisted (cheshmedjiev et al., 2010a) (fig. 1). environmental gradients and cyanoprokaryotes distribution the results of the pca, run on the environmental variables, total phytoplankton biomass (tbs) and total cyanoprokayotes biomass (tbc) of 61 wbs are shown in fig. 5. the cumulated relative inertia of the two first principal components reached 51.4 %. the main environmental gradient, associated with the first principal component, is determined by altitude and trophic status, with altitude (alt) positively correlated with secchi depth (sd), and negatively with tn and tp. unsurprisingly, tbs and tbc is correlated with the nutrient loading. the second principal component is mainly determined by temperature and depth (fig. 5). three groups of wbs are identified in this analysis: group 2 is constituted essentially by 3 closely situated coastal shallow wbs (namely atanasovsko ezero, vaya and mandra), which form the well-known geographical group bourgaski ezera) and group 3, formed by clear, oligotrophic alpine lakes and one high mountain reservoir (reservoir beli iskur, bezbozhsko ezero 1, nevenino ezero 1, gergiysko ezero 1, vlahinsko ezero 1, bunderishko ezero 9 and fig. 3. distribution of the most widespread genera of cyanoprokaryotes in the phytoplankton of bulgarian water bodies in the period 2000-2015. 0-40, number of water bodies in which species were found; *genus found in the samples containing cyanotoxins. no nco mm er cia l u se on ly m.p. stoyneva-gärtner et al.142 popovo ezero 2). the central group 1 is more heterogenous, comprising mostly mid-altitude wbs with varying depth and nutrient loading, but also some shallow eutrophic lakes (close to group 2, left side of the ordination). the distribution of the wbs in fig. 5 clearly shows the influence of geographic location on the wbs’ characteristics. the next redundancy analysis (fig. 6), in which the dependent variables are the biomass data on 24 cyanoprokaryote genera, showed that cyanoprokaryote assemblages responded significantly to the environmental conditions, with tp as the most influential variable, followed by sd, t and tn. four main groups of genera were identified: two (1 and 2) on the left side of the diagram, associated with eutrophic conditions in low altitude shallow lakes, swamps and reservoirs, one (group 3) of high altitude, clear lakes; and one (4) more heterogeneous group with predominance of high conductivity conditions. within this last are associated all rare colonial non-toxic species /rccl/ from genera like lemmermanniela, coelomoron, etc. and merismopedia (mainly m. tenuissima lemmermann) mrsp, non-colonial picoplankters of pcy group – pcpl) and spirulina and glaucospira species (spgl) with a single phormidium (phrm). in group 1 were microcystis mcrs; anabaena s.l. (mainly dolichospermum) – andl, aphanizomenon s.l. – aphn, cylindrospermopsis (in this case c. raciborskii (wołoszyńska) seenaya & subba raju only) clps, planktothrix – plnt, planktolyngbya s.l. (incl. limnolyngbya) – plnb, romeria rmrr, oscillatoria and borzia osbr. group 2 included anabaenopsis -anbs, pseudanabaena – psnb, plectonema – plct, leptolyngbya – lptb. group 3 contains aphanocapsa apns, chroococcus crcr, coelosphaerium clsp, snowella snwl, woronichinia wrnc) and synechocystis (sncs). cyanotoxins the analysis of published data on cyanotoxins revealed their presence in 16 wbs – reservoirs bistritsa, borovitsa, enitsa, kayabash 2, krushovitsa, mandra, pchelina, studena, studen kladenets, trakiets, vucha, valchovets and lakes vaya, durankulashko ezero, ezero momin brod and shablensko ezero (tab. 2, fig. 1). during the summer-autumn period, microcystins lr, la, rr, yr and similar to yr-type, nodularins, anatoxin-a and saxitoxins (from the decarbamoyl saxitoxin, gonyautotoxins ii, iii, b1, c1 and c2 group) in different concentrations were proved by high performance liguid chromatography (hplc, hplc-dad and/or hplc-ms), enzyme-linked immunosorbent assay (elisa) and in vitro cytotests (tab. 2). anatoxin was detected only once, in july 2006 in borovitsa reservoir by hplc (teneva et al., 2009). saxitoxins (stxs) were found in the reservoirs borovitsa, studen kladenets, trakiets and fig. 4. distribution of the number of species and genera of cyanoprokaryote phytoplankters in 115 bulgarian water bodies in the period 2000-2015. no nco mm er cia l u se on ly cyanoprokaryote blooms and cyanotoxins in bulgaria (2000-2015) 143 fig. 5. results of the principal component analysis on the environmental variables and phytoplankton/cyanoprokaryote biomass of 61 bulgarian water bodies (wbs); ordination on the two first components (cumulative inertia: 51.4%). numbers and names of the wbs as in tab. 1. t, water temperature; cond, electric conductivity; sd, secchi depth; tp, total phosphorus; tn, total nitrogen. fig. 6. results of the redundancy analysis on the environmental variables and cyanoprokaryote genera of 61 bulgarian water bodies (wbs); ordination on the two first components (cumulative inertia: 45.3%). numbers and names of the wbs as in tab. 1. ha, area; alt, altitude; t, water temperature; cond, electric conductivity; sd, secchi depth; tp, total phosphorus; tn, total nitrogen. no nco mm er cia l u se on ly m.p. stoyneva-gärtner et al.144 valchovets by teneva et al. (2009, 2010a, 2011) and stoyanov et al. (2012). their concentrations varied between 0.1 and 2.5 μg l–1, except in studen kladenets, where stxs were detected only by hplc peaks and authors suggest that their concentrations in the water were less than their mean lower detection limit by ridascreen™ assay – 0.01 μg l–1. in the same reservoir hplc peaks showed the presence of nodularins, but their amounts were given only as a total of microcystins/nodularins (mc/nod) detected by elisa (teneva et al., 2011). in the same way, mc/nod totals were provided for the reservoirs borovitsa, enitsa, kayabash 2, krushovitsa, studen kladenets and trakiets (teneva et al., 2009, 2010a, 2011, 2014; stoyanov et al., 2012). the types of microcystins were indicated in the publications for 10 wbs but only for 9 of them the values have been provided (tab. 2). according to pavlova et al. (2006, 2007, 2013a, 2014), pavlova (2007) and georgieva et al. (2015) the concentrations of the different microcystin types ranged as follows: lr – 0.1-8.3 μg l–1 in water samples and 14-292 μg g–1 in net samples, rr 0.1-12.7 μg l–1 in water samples and 2103.2 μg g–1 in net samples, yr – 0.13-5.5 μg l–1 in water samples and 0.8-71.8 μg g–1 in net samples. yr equivalent type was detected twice by hplc-dad in shablensko ezero (pavlova, 2007; pavlova et al., 2007) and la type was reported once from borovitsa by teneva et al. (2010b). negative results for cyanotoxins have been published by pavlova (2007), pavlova et al. (2007, 2013a, 2014, 2015), teneva et al. (2014) and georgieva et al. (2015) for 13 wbs and for some samples from 11 wbs with previously detected toxins (tab. 3). the distribution of wbs with toxins (16), the wbs with recorded blooms (14), wbs in which toxic species have been found (30) and of the not problematic wbs (54) in the 8 groups of their geographic location and vertical position is shown on fig. 7. according to the above cited references, 52 algae and some akinetes have been discovered in the water samples with detected toxins. among them 45 from 21 genera were identified at species or generic level. the information on the number of their findings in toxic samples, the number of wbs, from which these samples have been collected together with the range of species biomass exactly in these samples, types of toxins, methods used and sources are summarized in tab. 4. figs. 2 and 3 illustrate the distribution of these species and genera in bulgarian wbs during the studied period. fig. 7. distribution of water bodies (wb) with toxins, the wbs with recorded blooms, wbs in which toxic species have been found and of the not problematic wbs in the 8 phyla of their geographic location and vertical position (cv.i-lv.vi). colours are on conformity with colours on fig. 1, for the description of cv.i-lv.vi phyla see the text. no nco mm er cia l u se on ly cyanoprokaryote blooms and cyanotoxins in bulgaria (2000-2015) 145 ta b. 4 . c ya no pr ok ar yo te s fo un d in th e sa m pl es w ith d et ec te d to xi ns in b ul ga ri an w b s. ta xo n n f n w b b io m as s m ic ro cy st in s r r y r l a m c g o th er a n ta s t x s ou rc ra ng ee (m c ) to xi ns l r n o d m ic ro cy st is a er ug in os a (k üt z. ) k üt z. 1 5 9 x1. 4 x x x x x x pa vl ov a et a l. 20 06 , 2 01 5; t en ev a et a l. 20 09 , 2 01 0a . 2 01 1. 2 01 4; s to ya no v et a l. 20 12 a ph an iz om en on fo saq ua e r al fs e x b or n. e t f la h. 10 7 x2. 41 -d om x x x ? x x pa vl ov a et a l. 20 06 ; t en ev a et a l. 20 09 , 2 01 0a , b , 2 01 4; s to ya no v et a l. 20 12 m ic ro cy st is w es en be rg ii (k om .) k om . i n k on dr at ie va 1 0 5 0 .2 -1 1. 97 * x x x p av lo va e t a l. 20 06 ,2 01 4, 2 01 5 p se ud an ab ae na c at en at a l au te rb . 6 1 x -0 .0 5 x x x x t en ev a et a l. 20 11 m ic ro cy st is fl os -a qu ae (w itt r.) k ir ch n. 6 5 0. 01 -7 .2 x x x p av lo va e t a l. 20 06 , 2 01 5; s to ya no v et a l. 20 12 sy ne ch oc oc cu s el on ga tu s (n äg .) n äg . 6 3 0 .0 4d om x x x x x t en ev a et a l. 20 10 a, b, 2 01 1 d ol ic ho sp er m um a ffi ne (l em m .) w ac kl in e t a l. 4 2 x -0 .0 3 x x x x x x t en ev a et a l. 20 09 ; 2 01 1 d ol ic ho sp er m um s ch er em et ie vi i ( e le nk .) w ac kl in e t a l. 4 1 0 .0 20. 09 x x x t en ev a et a l. 20 11 a na ba en a sp . 4 4 0 .0 03 -8 .2 8 x x x p av lo va e t a l. 20 06 ; t en ev a et a l. 20 11 c hr oo co cc us m in ut us (k üt z. ) n äg . 4 2 0 .0 01 x x x x x x t en ev a et a l. 20 10 b, 2 01 1 sn ow el la la cu st ri s (c ho d. ) k om . e t h in d. 4 3 x -0 .0 2 x x x x ? x t en ev a et a l. 20 09 , 2 01 0a , b d ol ic ho sp er m um s pi ro id es (k le b. ) w ac kl in e t a l. 3 3 x 6 5. 33 x x x 3 x t en ev a et a l. 20 10 b, 2 01 4; s to ya no v et a l. 20 12 m er is m op ed ia te nu is si m a l em m . 3 2 x -0 .0 2 x x x p av lo va e t a l. 20 14 . 2 01 5 m ic ro cy st is n at an s l em m . e x sk uj a 3 3 0. 05 *1. 47 * x x x pa vl ov a et a l. 20 06 m ic ro cy st is sp . 3 2 0. 1* -n ot in d. x p av lo va e t a l. 20 15 d ol ic ho sp er m um fl os -a gu ae (b ré b. e x b or n. e t f la h. ) w ac kl in e t a l. 2 1 x x x x ? x te ne va e t a l. 20 10 b d ol ic ho sp er m um s ol ita ri um (k le b. ) w ac kl in e t a l. 2 1 0. 10. 18 x t en ev a et a l. 20 11 tr ic ho rm us v ar ia bi lis (k üt z. e x b or n. e t f la h. ) k om . e t a na gn . 2 1 x x x x t en ev a et a l. 20 11 p la nk to th ri x ag ar dh ii (g om .) a na gn . e t k om . 2 2 2 .3 210 .5 x s to ya no v et a l. 20 12 p ho rm id iu m s p. 2 1 x x te ne va e t a l. 20 11 sp ir ul in a m aj or k üt z. e x g om . 2 2 x -0 .0 3 x x x t en ev a et a l. 20 11 ; p av lo va e t a l. 20 15 a ph an iz om en on s p. ju v. 1 1 0 .0 98 x pa vl ov a et a l. 20 14 a ph an oc ap sa d el ic at is si m a w . e t g .s . w es t 1 1 0 .6 x x pa vl ov a et a l. 20 15 a ph an oc ap sa g re vi lle i ( b er k. ) r ab en h. 1 1 0 .6 2* x p av lo va e t a l. 20 06 a ph an oc ap sa s pp . 1 1 0 .0 04 * x x g eo rg ie va e t a l. 20 15 c hr oo co cc us a ph an oc ap so id es s ku ja 1 1 x x p av lo va e t a l. 20 15 c hr oo co cc us d is pe rs us (k ei ss l.) l em m . 1 1 0. 02 t en ev a et a l. 20 11 c hr oo co cc us s p. 1 1 0 .0 01 x x x pa vl ov a et a l. 20 14 c yl in dr os pe rm op si s ra ci bo rs ki i ( w oł .) se en ay a et s ub ba r aj u 1 1 x x s to ya no v et a l. 20 12 le pt ol yn gb ya fo ve ol ar um (r ab en h. e x g om .) a na gn . e t k om . 1 1 0. 01 1 x x x pa vl ov a et a l. 20 14 li m no ra ph is h ie ro no m us ii (l em m .) k om . e t a l. 1 1 x x x te ne va e t a l. 20 11 li m no th ri x sp . 1 1 4 9. 93 * x x x pa vl ov a et a l. 20 06 m er is m op ed ia g la uc a (e hr .) k üt z. 1 1 x x x pa vl ov a et a l. 20 15 m er is m op ed ia h ya lin a (e hr .) k üt z. 1 1 0 .0 08 x x x pa vl ov a et a l. 20 14 to b e co nt in ue d on n ex t p ag e no nco mm er cia l u se on ly m.p. stoyneva-gärtner et al.146 discussion the analysis of data on phytoplankton species composition revealed that, in spite of being quite heterogeneous, they are generally available for 111 from 115 purposively studied wbs included in tab. 1, except concrete data for the reservoirs stoychovtsi, karaisen, luzhenska bara and aleksandrovo. however, the occurrence of cyanoprokaryotes in the last reservoir was documented as “presence of toxic species” (cheshmedjiev et al., 2010a; fig. 1). in all samples proceeded by us, cyanoprokaryotes were not found in 11 alpine lakes (bunderishko ezero 9, musalensko ezero 3, ezero bliznaka, ezero bubreka, ezero okoto, ezero sulzata, kremensko ezero 2, marichino ezero 2, marichino ezero 3, popovo ezero 2 and vlahinsko ezero 1) and in the occasionally sampled reservoirs yarlovets and zhernov. therefore, we argue that 99 wbs contain different number and amount of cyanoprokaryotes. with the total of 210 taxa found, they represent only 3.8% of the algal biodiversity of the country estimated as ca. 5,500 taxa by stoyneva (2014), but comprise 36.4% from all 576 cyanoprokaryotes (stoyneva et al., 2016). coccal cyanoprokaryotes were the richest group both in species (85) and genera (31), while the number of heterocytous species and genera was the lowest. this result is on general conformity with the shallow and euto hypertrophic character of most studied wbs, in which, according to our unpublished results, nitrogen was not a limiting factor (n/p ranged between 7 and 110). similar conclusion could be made according to the data in the publications by teneva et al., 2010a; belkinova et al., 2014; pavlova et al., 2015; stoyanov et al., 2016, etc. the highest taxonomic diversity, expressed as number of species and genera, was detected in two wbs: blato srebarna (85/35) and vaya (63/30) (fig. 4). due to their high importance for conservation of rare and threatened species of national, european and global significance (michev and stoyneva, 2007; vassilev et al., 2013), these two shallow lakes have been intensively studied before and during the analyzed period, and always have been outlined for their rich algal diversity (michev et al., 1998; stoyneva, 1998a, 1998b, 2003, 2014, 2015; georgiev, 2012; dimitrova et al., 2014a, 2014b). however, the same authors outlined the negative trends in their development with enhanced eutrophication, cyanoprokaryote blooms (incl. the rare equilibrium states dominated by cyanoprokaryotes) formed by potentially toxic species. the negative effects of increased eutrophication due to long-term cage fish farming were stressed also for the mountain reservoirs dospat and kardzhali, where among the newly appeared group of cyanoprokaryotes the harmful aphanizomenon flos-aquae, dolichospermum spiroides and planktothrix rubescens (de candolle ex gomont) anagnostidis et komárek participated in the dominant complexes (dochin and stoyneva, 2014, 2015; dochin 2015).ta b. 4 .c on tin ue d fr om p re vi ou s pa ge . ta xo n n f n w b b io m as s m ic ro cy st in s r r y r l a m c g o th er a n ta s t x s ou rc ra ng ee (m c ) to xi ns l r n o d m ic ro cy st is b ot ry s te il. 1 1 0 .6 x x x pa vl ov a et a l. 20 15 m ic ro cy st is fi rm a (k üt z. ) s ch m id le 1 1 0 .0 4* p av lo va e t a l. 20 06 m ic ro cy st is p ul ve re a (w oo d) f or ti in d e to ni 1 1 d om x x x ? x te ne va e t a l. 20 10 b o sc ill at or ia a nn ae v an g oo r 1 1 x x x te ne va e t a l. 20 11 o sc ill at or ia s p. 1 1 0 .0 3 x te ne va e t a l. 20 09 p la nk to th ri x co m pr es sa (ü te rm oh l) a na gn . e t k om . 1 1 0 .0 7 x te ne va e t a l. 20 11 p se ud an ab ae na li m ne tic a (l em m .) k om . 1 1 0 .1 x t en ev a et a l. 20 11 p se ud an ab ae na m uc ic ol a (n au m an n et h ub .-p es t.) b ou rr . 1 1 0. 11 8 x x x pa vl ov a et a l. 20 14 r ap hi di op si s m ed ite rr an ea sk uj a 1 1 0. 00 10. 11 8 x p av lo va e t a l. 20 14 w or on ic hi ni a na eg el ia na (u ng .) e le nk . 1 1 0. 01 * x x x pa vl ov a et a l. 20 06 w or on ic hi ni a sp . 1 1 0. 48 * x x x pa vl ov a et a l. 20 06 u ni de nt if ie d [s ol ita ry c el ls . a ki ne te s. fi la m en ts -2 . c oc ca l c ol on ia l 1] 9 6 x22 .5 4* x x x p av lo va e t a l. 20 06 , 2 01 4, 2 01 5 n f, nu m be r o f f in di ng s/ sa m pl es ; n w b , n um be r o f w b s in w hi ch th e sp ec ie s w as re co rd ed ; b io m as s ra ng e, re la te d to th e sp ec ie s in th e “t ox ic ” sa m pl es , m c g, in di ca te d on ly a s m ic ro cy st in s in th e pu bl ic at io ns (w ith ou t t he ir e xa ct ty pe s) ; x , p re se nc e; d om , d om in an t s pe ci es /d om in an ce ; * bi om as s es tim at ed fo r t hi s st ud y; n ot in d. , n ot in di ca te d in th e pu bl ic at io ns . no nco mm er cia l u se on ly cyanoprokaryote blooms and cyanotoxins in bulgaria (2000-2015) 147 the analysis of the distribution of species and genera points to 2 species (aphanizomenon flos-aquae and cylindrospermopsis raciborskii) and 11 genera as being the most widespread in the country (planktothrix, cylindrospermopsis, dolichospermum, chroococcus, microcystis, pseudanabaena, anabaena, aphanocapsa, planktolyngbya and aphanizomenon; figs. 2 and 3). except for chroococcus, aphanocapsa and planktolyngbya they have been commonly pointed as principally responsible for forming blooms (e.g. oliver and ganf, 2002) mainly due to presence of gas vesicles. among the total of 69 genera found, at least 29 are known as cyanotoxin producers (metcalf and codd, 2012; pettersson and pozdnyakov 2013). if the genera chrysosporum and sphaerospermopsis, which relatively recently have been separated from anabaena (komárek, 2013 and references therein), are considered, this number will increase to 31, or 45% from all found in the country. the frequency of findings of all these potentially toxic genera in combination with the spreading of cyanoprokaryote blooms (fig. 1) are logically related with the findings of different types and amounts of cyanotoxins in bulgarian wbs (tab. 2). the results of multivariate analysis show the consistency of the wbs classification based on geographic position and altitude (as first steps proposed in the hierarchical classification of bulgarian wetlands by michev and stoyneva, 2007), along with in-lake characteristics as depth and nutrient content. total phytoplankton biomass, total cyanoprokaryote biomass, as well as cyanoprokaryote assemblages showed a strong response to the environmental variables, with an expected major influence of tp. the redundancy analysis identified four groups of genera in relation to wbs, among which groups 3 and 4 are more heterogenous in comparison with groups 1 and 2. group 3 contains realtively small colonial embedded in mucilage coccal genera (aphanocapsa, chroooccus, coelosphaerium, snowella, woronichinia) and non-colonial coccal synechocystis which are classical phytoplankters in various types of wbs. despite their general ubiquity and abundance in a wide spectrum of trophic conditions, these non-bloom formers are not a well-known group, particularly in relation to their ecology (stockner et al., 2002). in our opinion, the presence of all these small coccal genera in group 3 reflects well the heterogenous character of its wbs in terms of geographic location, morphometry and way of use, but united by the prevalence of mesoto oligotrophic conditions. within the group 4 with predominantly high conductivity conditions, characterized by extremely low tbc, are associated diverse and most rarely (in this study) distributed smallest coccal cyanoprokaryotes (all rare colonial non-toxic species, merismopedia, non-colonial picoplankters of pcy group), very thin filamentous rare spirulina and glaucospira species with a single (and most probably benthic) phormidium. all of them are generally considered as non-blooming cyanoprokaryotes (stockner et al., 2002). by contrast, most genera well known for their ability to form water blooms due to presence of gas vesicles (oliver and ganf, 2002) were in group 1: microcystis; anabaena s.l. (mainly dolichospermum), aphanizomenon s.l., cylindrospermopsis and planktothrix. the other genera in group 1 are filamentous without gas vesicles: classical planktonic thin planktolyngbya s.l., generally short-celled fine romeria and occasionally distributed in this data set short-celled oscillatoria and borzia. as it could be seen, besides the very rare and therefore non-representative osbr, the “exception” from this filamentous group 1 is the coccal colonial microcystis. however, the “filamentous” character of another participant in group 1 romeria is still not proved and its coccal nature is in discussion (komárek and anagnostidis, 2005). common feature of the strangers (microcystis, planktolyngbya and romeria) in group 1 is the presence of well-developed homogenous colourless mucilage. the irregular presence of gas vesicles in combination with wide, but fine mucilage envelopes is known also for anabaenopsis and pseudanabaena, which are representatives of group 2. the group includes also the thin (1-2 µm) filamentous rarely branched plectonema and straight filamentous leptolyngbya representatives, most of which have not been determined at species level due to their small dimensions and problematic taxonomy (komárek and anagnostidis, 2005). the stranger in the group is the heterocytous and capable of bloom-forming anabaenopsis. however, a capability for nitrogen fixation was proved for some leptolyngbya species also (e.g. stal, 2012). according to our recent knowledge, very few data confirm the toxic abilities of the genera in group 2 (except for anabaenopsis milleri voronichin, which was not found in our studies), which comprises coastal and inland lowland lakes and reservoirs with generally hypertrophic character. by contrast, the genera from group 1, and its core representatives (aphanizomenon, cylindrospermopsis, microcystis and planktothrix) in particular were repeatedly pointed as bloom-forming and toxin producing genera related mainly with eutrophic conditions and were also the key players in toxic samples, detected in bulgarian waters (fig. 3; tab. 2). they were found among the genera with broadest distribution in the country (fig. 3). the common traits of the taxa belonging to this group are the affinity for eutrophic conditions and the presence of gas vesicles, which provide buoyancy control of filament and colonies, allowing efficient vertical migration in the water column of relatively shallow lakes which present alternation of periods of stratification and mixing (reynolds, 2006). accordingly, the wbs of group 1 are relatively small, inland lowland to low-mountain, euto hypertrophic reservoirs. the samples with detected cyanotoxins were from 16 no nco mm er cia l u se on ly m.p. stoyneva-gärtner et al.148 wbs (tab. 2) and contained 44 cyanoprokaryotes identified to species or genus level (tab. 4). their distribution in bulgaria, with a few exceptions, is relatively broad (figs. 2 and 3). the results from this study confirm the trend for their fast spread in the country outlined especially in relation to the invasive species like cylindrospermopsis raciborskii (stoyneva, 2015; kokocinski et al., in press). most of these 44 species and 21 genera have been repeatedly reported as real or potential toxin producers but for some of them discussions are still running (e.g. for microcystis species as summarized in šejnohová and maršálek, 2012). it is out of scope of this study to go deep in the contradictory opinions, which have to be interpreted with caution, moreover for details in relation with findings in bulgaria we can refer to our previous papers (pavlova et al., 2006, 2014, 2015). summarizing the results from this study, we could outline that exactly one of the most often discussed species m. wesenbergii was the most often recorded species in toxic samples (tab. 4). in the same time, pseudanabaena mucicola (naumann et huber-pestalozzi) bourrelly, which is a relatively regular endophyte in m. wesenbergii, was reported only once as associated with cyanotoxin detection. some of the species in tab. 4 are identified only at genus level, or are known for problematic identification, and therefore discussion on them is not relevant without proper documentation, which is not available in the published papers. it has to be noted also, that other cyanoprokaryotes, known for their potential to produce toxins, were found in the country. for example, nodularia spumigena mertens ex bornet ex flahault was reported from three temporary wbs on belene island in the danube (beshkova and botev, 2004), but the study was not supplied with toxin analysis. obviously, aimed studies on toxin groups other than microcystins, combined with permanent toxin monitoring will reveal more objective picture of presence/absence of different cyanotoxins and species related with them. a cooccurrence of other algae could also have an effect, and even provide a source, for some of the toxins detected (e.g. pyrrhophytes for saxitoxins). since long time it is well-known that cyanoprokaryote dominance (and blooms in particular) is strongly affected by complex interactions between lake morphometry, water temperature, underwater light availability, nutrient supply and total food-web structure and can not be viewed independently from all other members of phytoplankton (harris, 1986; reynolds, 1987, 2006; dokulil and teubner, 2000 among the many others). therefore, we would like to outline the necessity for providing data in future publications on the total species composition for the samples and wbs with detected cyanotoxins or harmful blooms. in the temperate country like bulgaria it is not surprising that cyanotoxins were detected in the water samples from summer-autumn periods, when the abundance of cyanoprokaryotes is normally the highest for the year. however, the comparison of tabs. 2 and 3 clearly shows that occurrence of toxins is intermittent in bulgarian wbs. conclusions about the correlations between finding of cyanotoxins and relevant species or total sample biomass are hardly possible due to lack of concrete published data for all cases. the only statement which could be done is that cyanotoxins were detected within a broad range of algal abundance (tabs. 2 and 4). in some cases, dolichospermum spiroides (klebahn) wacklin et al., limnothrix sp., microcystis wesenbergii, planktothrix agardhii (gomont) anagnostidis et komárek, anabaena sp. and some unidentified algae had high biomass values in samples with detected toxins (tab. 4). but toxins were detected even when species were found in low, even negligible, concentrations (the cases of borovitsa, valchovets, studena, etc. in tab. 2). the topic was discussed by pavlova et al. (2006) and teneva et al. (2009), who supposed that such cyanotoxin quantities were due to blooms, which occurred in the wbs before the sampling. therefore, it is necessary to outline again that finding of cyanotoxins in waters is not always related with obvious blooms. moreover, the term bloom and its lower borders remained poorly defined (oliver and ganf, 2002). however, the result obtained is important for rising the public awareness of the cyanoblooms, which could have long consecutive effect on the ecosystem health and humans. in spite of heterogenous way of representing the results on cyanotoxin amounts, it could be stated that in most cases, and in the drinking-water reservoirs in particular, the concentration of microcystin lr is lower than who’s (1998) limit of 1 μg l–1 (tab. 2). the general amount and types of microcystins and other cyanotoxins are on conformity with the results published for the closest neighboring countries macedonia, serbia, romania, turkey and greece, as it was outlined by pavlova et al. (2015). however, the fact of their findings in 16 wbs (among which 3 are important drinking-water reservoirs, 2 are water supply reservoirs, 3 are used for irrigation and 8 are recreational sites for sports and fishing) on the background of more recorded cyanoprokaryote blooms and broad spread of toxic species in bulgaria (figs. 1, 2 and 3) is strong enough to alarm both scientists and responsible authorities at national level. moreover, the analyses of all data clearly show that during the last 15 years cyanoprokaryotes were not restricted in distribution in the lowlands and plains but started to invade wbs with higher altitudes, i.e. wbs situated in kettles and low mountains (fig. 7). the outlining of the necessity of stronger recognition of the problem in the country and rising of the public awareness made here is not the first, just the opposite – almost each of the papers on the topic published in the analyzed period had this statement as a main conclusion since there is no doubt that the peculiar and otherwise fascinating group of cyanoprokaryotes is a no nco mm er cia l u se on ly cyanoprokaryote blooms and cyanotoxins in bulgaria (2000-2015) 149 real hazardous factor for human and aquatic ecosystem health in bulgaria. conclusions all results from studies carried in summerautumn periods of the last 15 years (2000-2015) in 120 bulgarian wbs different in location, morphometry and trophic status (incl. drinking-water reservoirs, reservoirs for water supply, irrigation and energy production reservoirs, recreational lakes and sites of nature conservation importance), gathered from 35 publications, showed that in 30 of them toxic species were found. cyanoprokaryote blooms were recorded in 14 wbs and in 16 cyanotoxins (microcystins, nodularins and saxitoxins) were detected, and cyanoprokaryote diversity was quite high (210 taxa of 60 genera). toxin concentration ranged between 0.1 and 26.5 µg l–1 in water samples and between 10.9 and 1070 µg g–1 (d.w.) in concentrated (net) samples. despite the fact that microcystins were not found in all studied wbs and that the recorded levels were still lower in comparison with some other european countries and with who’s threshold for microcystin lr, the fact of cyanotoxin detection in 3 drinking-water reservoirs and cyanoprokaryote occurrence in low mountain wbs could serve as an alert for the need of recognition of cyanotoxins as a new health risk factor in the country. therefore, permanent monitoring with identification of toxins in wbs at risk and activities for limitation and control of toxic blooms are urgently needed, in combination with increase of the attention to the effects of cyanotoxins on both human health and health of aquatic ecosystems in bulgaria. we also stress the need for a more comprehensive monitoring of the problematic wbs and of their watershed – including all key environmental variables (hydrology, nutrient loading and meteorology) and detailed phytoplankton surveys in order to improve water quality management and identify the measures to be taken to reduce the risks associated with cyanoprokaryote blooms. acknowledgments the authors would like to acknowledge the european cooperation in science and technology, cost action es 1105 “cyanocostcyanobacterial blooms and toxins in water resources: occurrence, impacts and management” and bulgarian national supporting financing by the scientific fund of the ministry of education and science (projects dkost 01/2-11.08.2016 and dkost 10/2-19.08.2016) for adding value to this study through networking and knowledge sharing with european experts and researchers in the field. a part of the results, taken from stoyneva (2014), was obtained during emerge project of ec framework v, mews project on srebarna lake monitoring (19942003), project of world bank and zeleni balkani for management plan of pomorie wetland, monitoring studies of bulgarian lakes of ecotan eood commissioned by the eea of mews of bulgaria and studies on chosen oligotrophic lakes in pirin mts, commissioned by pirin national park. the authors are thankful to the colleagues assoc. prof t. michev, assist. b. michev, mr n. mihov, mr r. rusev, mr. p. simeonov, assoc. prof. dr p. zhelev, dr a. asenov, main assist. dr p. ivanov and assoc. prof. dr i. traykov for collecting samples and environmental data out of the frame of the above mentioned projects. special thanks are due to mrs. a. lazarova from the national health centre for analyzing of tp and tn in water samples, and to two anonymous reviewers for helpling improve the manuscript. references belkinova d, gecheva g (eds.), 2013. [biological analysis and ecological assessment of the types of surface waters in bulgaria].[book in bulgarian]. plovdiv university “paisiy hilendarski” publishing, plovdiv: 201 pp. belkinova d, mladenov r, dimitrova-dyulgerova i, cheshmedjiev s, angelova i, 2007. phytoplankton research in kurdzhali reservoir. phytol. balc. 13:47-52. belkinova d, padisák j, gecheva g, cheshmedjiev s, 2014. phytoplankton based assessment of ecological status of bulgarian lakes and comparison of metrics within the water framework directive. appl. ecol. environ. res. 12:83-103. beshkova m, kalchev r, vasilev v, 2012. taxonomical and functional structure, species diversity and abundance of phytoplankton assemblages of the srebarna lake in relation to changes of ecological conditions, p. 39-56. in: y. uzunov, b.b. georgiev, e. varadinoiva, n. ivanova, l. pehlivanov and v. vasilev (eds.), ecosystems of the biosphere reserve srebarna lake. professor marin drinov academic publishing house. beshkova m, kalchev rk, vassilev v, tsvetkova rl, 2008a. changes of the phytoplankton abundance and structure in the biosphere reserve srebarna (northeastern bulgaria) in relation to some envronmental variables. acta zool. bulg. suppl. 2:166-174. beshkova mb, botev is, 2004. phytoplankton community structure of three temporary wetlands on belene island (bulgarian sector of the danube river). phytol. balc. 10:11-19. beshkova mb, kalchev rk, kalcheva hv, 2008b. phytoplankton and bacterioplankton in three reservoirs (north east bulgaria) accepted as a potential referent sites according the water framework directive of eu. acta zool.bulg. suppl. 2:155-164. beshkova mb, kalchev rk, kalcheva hv, 2014. phytoplankton in the zhrebchevo reservoir (central bulgaria) before and after invasion of dreissena polymorpha (mollusca: bivalvia). acta zool. bulg. 66:399-409. no nco mm er cia l u se on ly m.p. stoyneva-gärtner et al.150 callieri c, stockner jg, cronberg g, 2012. freshwater picocyanobacteria: single cells, microcolonies and colonial forms, p. 229-270. in: b.a. whitton (ed.), the ecology of cyanobacteria ii. their diversity in space and time. springer. carmichael ww, 1994. the toxins of cyanobacteria. sci. am. 270:64-72. cen en 15204, 2006. water quality guidance standard on the enumeration of phytoplankton using inverted microscopy (utermöhl technique): 42 pp. cheshmedjiev s, belkinova d, mladenov r, dimitrova-dyulgerova i, gecheva g, 2010a. phytoplankton based assessment of the ecological status and ecological potential of lake types in bulgaria. biotechnol. biotec. eq. 24:14-25. cheshmedjiev s, gecheva g, belkinova d, varadinova e, dimitrova-dyulgerova i, mladenov r, soufi r, pavlova m, pehlivanov l, 2013. assessment of ecological status and preliminary results on reference conditions in alpine glacial lakes (bulgaria) a contribution to the implementation of the water framework directive. biotechnol. biotec. eq. 27:3522-3528. cheshmedjiev s, karagiozova t, michailov m, valev v, 2010b. revision of river and lake typology in bulgaria within ecoregion 12 (pontic province) and ecoregion 7 (eastern balkans) according to the water framework directive. ecologia balkanica 2:75-96. chorus i, bartram j (eds.), 1999. toxic cyanobacteria in water. a guide to their public health consequences, monitoring and management. who, publ. e & fn spon, london: 416 pp. codd ga, 1994. cyanobacterial (blue-green algal) toxins in marine and estuarine waters. scott. assoc. mar. sci. newslett. 9:7. codd ga, 1995. cyanobacterial toxins: occurrence, properties and biological significance. water sci. technol. 32: 49-156. codd ga, azevedo smfo, bagchi sn, burch md, carmichael ww, harding wr, kaya k, utkilen hc, 2005b. cyanonet: a global network for cyanobacterial bloom and toxin risk management. initial situation assessment and recommendations, technical documents in hydrology 76, unesco, paris: 138 pp. codd ga, bell sg, kaya k, ward cj, beattie ka, metcalf js, 1999. cyanobacterial toxins, exposure routes and human health. eur. j. phycol. 34:405-415. codd ga, morrison lfm, metcalf js, 2005a. cyanobacterial toxins: risk management for health protection. toxicol. appl. pharm. 203:264-272. dimitrova re, nenova ep, uzunov ba, shihiniova md, stoyneva mp, 2014a. phytoplankton abundance and structural parameters of the critically endangered protected area vaya lake (bulgaria). biotechnol. biotec. eq. 28:871-877. dimitrova re, nenova ep, uzunov ba, shihiniova md, stoyneva mp, 2014b. phytoplankton composition of vaya lake (2004-2006). bulg. j. agric. sci. 20. suppl. 1:165-172. dochin kt, 2015. [seasonal dynamics and species composition of the phytoplankton in the reservoirs kardzhali and dospat].[phd thesis in bulgarian], sofia university “st kliment ohridski”, faculty of biology. dochin kt, stoyneva mp, 2014. effect of long-term cage fishfarming on the phytoplankton biodiversity in two large bulgarian reservoirs. ber. nat.-med. verein innsbruck 99:49-96. dochin kt, stoyneva mp, 2015. phytoplankton of the reservoir “dospat” (rodopi mts, bulgaria) as indicator of negative trend in reservoir development due to long-term cage fish farming. ann. sof. univ. fac. biol. book 2 botany 99:47-60. dokulil mt, teubner k, 2000. cyanobacterial dominance in lakes. in: p.b. hamilton, h. kling and m.t. dokulil (eds.), cyanoprokaryotes and chlorophytes across lake trophic states. hydrobiologia 438:1-12. european commission, 2000. directive of the european parliament and of the council 2000/60/ec establishing a framework for community action in the field of water policy. official journal l 327:1-72. georgiev b, 2012. biodiversity of srebarna biosphere reserve: an overview, p. 13-24. in: y. uzunov, b.b. georgiev, e. varadinoiva, n. ivanova, l. pehlivanov and v. vasilev (eds.), ecosystems of the biosphere reserve srebarna lake. professor marin drinov academic publishing house. georgieva v, pavlova v, bratanova zl, 2015. hygienic assessment of the water reservoirs “studena”, “bistritsa” and “pchelina”, based on performed hydrobiological analysis and determination of some toxins. bulg. j. public health 7:3-13. graham le, graham jm, wilcox lw, 2009. algae. benjamin cummings, san francisco: 616 pp. griffiths hi, kryštufek b, reed jm, 2004. balkan biodiversity. pattern and process in the european hotspot. kluwer, dordrecht: 357 pp. guiry md, guiry gm, 2016. algaebase. world-wide electronic publication, national university of ireland, galway. http://www.algaebase.org harris gp, 1986. phytoplankton ecology. structure, function and fluctuation. chapmann and hall, london, new york: 384 pp. huisman j, matthijs hcp, visser pm, 2005. harmful cyanobacteria. springer, the netherlands, dordrecht: 241 pp. kalchev r, botev i, hristozova m, naidenow w, raikowapetrova g, stoyneva m, temniskova-topalova d, trichkova t, 2004. ecological relations and temporal changes in the pelagial of the high mountain lakes in the rila mountains (bulgaria). j. limnol. 63:90-100. kokociński m, akçaalan r, salmaso n, stoyneva m, sukenik a, 2016. expansion of alien/invasive cyanobacteria. in: j. meriluoto, g.a. codd and l. spoof (eds.), handbook of cyanobacterial monitoring and cyanotoxin analysis, first edition. john wiley & sons (in press). komárek j, 2013. cyanoprokaryota. 3rd part: heterocytous genera. in: b. büdel, l. krienitz, g. gärtner, and m. schagerl (eds.), süßwasserflora von mitteleuropa, 19(3), elsevier: 1130 pp. komárek j, anagnostidis k, 1999. cyanoprokaryota. 1 teil: chroococcales. in: h. ettl, g. gärtner, h. heynig and d. mollenhauer (eds.), süßwasserflora von mitteleuropa, 19(1), g. fischer, jena: 549 pp. komárek j, anagnostidis k, 2005. cyanoprokaryota. 2nd part: oscillatoriales. in: b. büdel, l. krienitz, g. gärtner and m. schagerl (eds.), süßwasserflora von mitteleuropa, 19(2), elsevier: 758 pp. komárek j, kastovsky j, jezberová j, 2011. phylogenetic and taxonomic delimitation of the cyanobacterial genus aphanothece and description of anathece gen. nov. eur. j. phycol. 46:315-326. legendre p, gallagher ed, 2001. ecologically meaningful transformations for ordination of species data. oecologia 129:271-280. no nco mm er cia l u se on ly cyanoprokaryote blooms and cyanotoxins in bulgaria (2000-2015) 151 maršálek b, bláha l, hindák f, 2000. review of toxicity of cyanobacteria in slovakia. biologia 55:645-652. mcneil j, barrie fr, bruck wr, demoulin v, greuter v, hawksworth dw, herendeen ps, knapp s, marhold k, prado j, prud’homme van reine wf, smith gf, wiersema jh, turland nj (eds.), 2012. international code of the nomenclature for algae, fungi and plants (melbourne code). regnum vegetabile 154. koeltz scientific books: 240 pp. meriluoto j, codd ga (eds.), 2005. toxic: cyanobacterial monitoring and cyanotoxin cnalysis. åbo akademi university press, turku: 149 pp. metcalf js, codd ga, 2012. cyanotoxins, p. 651-675. in b.a. whitton (ed.), ecology of cyanobacteria ii: their diversity in space and time, springer. michev tm, georgiev bb, petrova av, stoyneva mp (eds.), 1998. biodiversity of the srebarna biosphere reserve. checklist and bibliography. co-publ. context & pensoft, sofia: 130 pp. michev tm, stoyneva mp (eds.), 2007. inventory of bulgarian wetlands and their biodiversity. part 1: non-lotic wetlands. publishing house elsi-m, sofia: 364 pp. michev tm, stoyneva mp, 2005. red list of bulgarian wetlands: conception, creation and application. ann. sofia univ. 96: 71-76. mischke u, carvalho l, mcdonald c, skjelbred b, solheim al, phillips g, de hoyos c, borics g, moe j, pahissa j, 2011. wiser: deliverable d3.1-2: report on phytoplankton bloom metrics. project co-funded by the european commission within the seventh framework programme (20072013) dissemination level: 48 pp. oliver rl, ganf gg, 2002. freshwater blooms, p. 149-194. in: b. a. whitton and m. potts (eds.), the ecology of cyanobacteria. their diveristy in time and space. kluwer. pavlova v, 2007. hygiene and analytical aspects of microcystins occurrence in surface water. [ph.d. thesis in bulgarian], national center of public health protection, sofia. pavlova v, babica p, todorova d, bratanova z, maršalek b, 2006. contamination of some reservoirs and lakes in republic of bulgaria by microcystins. acta hydrochim. hydrobiol. 34:437-441. pavlova v, stoyneva m, bratanova z, 2013a. cyanoprokaryotes (cyanobacteria) and cyanotoxins in some bulgarian reservoirs. j. balkan ecol. 16:257-260. pavlova v, stoyneva m, bratanova z, karadjova i, 2013b. [cyanoprokaryotes (cyanobacteria) and cyanotoxins – health aspects and safety requirements for air transportation, p. 96100].[article in bulgarian with english abstract]. in: proceedings scientific conference “actual problems of safety”, 16-18.10.2013, publ. complex of the nvu “vasil levski. pavlova v, stoyneva m, georgieva v, donchev d, spoof l, meriluoto j, bratanova z, karadjova i, 2014. new records of microcystins in some bulgarian water bodies of health and conservational importance. j. water res. protect. 6: 446-453. pavlova v, stoyneva m, babica p, kohoutek j, bratanova z, 2007. microcystins contamination and cyanoprokaryote blooms in some coastal bulgarian wetlands, p. 221-226. in: conf. preprint book bulaqua 2007, sofia. pavlova v, stoyneva-gärtner m, uzunov b, bratanova z, lazarova a, karadjova i, 2015. microcystins -lr, -yr and -rr in six bulgarian water bodies of health and conservational importance (2012-2014). j. water res. protect. 7:1375-1386. peev d, petrova a, anchev m, temnsikova d, denchev cm, ganeva a, gussev c, vladimirov v (eds.), 2013. red data book of the republic of bulgaria. vol. 1. plants and fungi. institute of biodiversity and ecosystem research of bulgarian academy of sciences and ministry of environment and waters of bulgaria & moew: 525 pp. pettersson lh, pozdnyakov d, 2013. monitoring of harmful algal blooms. springer, dordrecht: 307 pp. reynolds cs, 1987. cyanobacterial water blooms. adv. bot. res. 13:67-143. reynolds cs, 1996. plant life of the pelagic. verh internat verein theor angew limnol. 26:97-113. reynolds cs, 2006. ecology of phytoplankton. cambridge university press, cambridge: 535 pp. rott e, 1981. some results from phytoplankton counting intercalibration. schweiz. z. hydrol. 43:34-62. schopf jw, 2012. the fossil record of cyanobacteria, p. 15-38. in: in: b.a. whitton (ed.), the ecology of cyanobacteria ii. their diversity in space and time. springer. šejnohová l, marsalek b, 2012. microcystis, p. 195-228. in: in: b.a. whitton (ed.), the ecology of cyanobacteria ii. their diversity in space and time. springer. sommer u, padisák j, 1993. appendix 1. identification of equilibrium. in: u. sommer, j. padisák, c.s. reynolds and p. juhász-nagy (eds.), hutchinson’s heritage: the diversity-disturbance relationship in phytoplankton. hydrobiologia 249:5-6. stal lj, 2012. cyanobacterial mats and stromatolites, p. 65-126. in: b.a. whitton (ed.), the ecology of cyanobacteria ii. their diversity in space and time. springer. state order n 4/14.09.2012, 2013. for characterization of the surface waters of the minister of environment and waters. state gazette 22. stockner jg, callieri c, cronberg g, 2002. picoplankton and other non-bloom forming cyanobacteria in lakes, p. 195231. in: b.a. whitton (ed.), the ecology of cyanobacteria ii. their diversity in space and time. springer. stoyanov ps, 2014. [variability, taxonomy and ecology of some filamentous blue-green algae (cyanoprokaryota)].[phd thesis in bulgarian], plovdiv university “paysiy hilendarski”, plovdiv. stoyanov p, belkinova d, mladenov r, teneva i, 2012. [analysis of the water in the reservoirs krushovitsa, enitsa and valchovets (northern bulgaria) for presence of cyanotoxins, p. 237-249].[article in bulgarian with english abstract]. in: pu “p. hilendarski”, jubilee proceedings “biological sciences for a better future”. stoyanov p, teneva i, mladenov r, belkinova d, 2013. diversity and ecology of the phytoplankton of filamentous bluegreen algae (cyanoprokaryota, nostocales) in bulgarian standing waters. ecologia balkanica 5:1-6. stoyanov p, teneva i, mladenov r, belkinova d, 2016. filamentous cyanoprokaryotes (cyanoprokaryota/ cyanobacteria) in standing waters of bulgaria: diversity and ecology. j. biosci biotechnol. 5:19-28. stoyneva m, 1998a. algae, p. 10-37. in: t.m. michev, b.b. georgiev, a.v. petrova and m. p. stoyneva (eds.), biodiverno nco mm er cia l u se on ly m.p. stoyneva-gärtner et al.152 sity of the srebarna biosphere reserve. checklist and bibliography. co-publ. context & pensoft. stoyneva mp, 1998b. development of the phytoplankton of the shallow srebarna lake (north-eastern bulgaria) across the trophic gradient. hydrobiologia 369/370:259-367. stoyneva mp, 2003. steady-state phytoplankton assemblages in shallow bulgarian wetlands. hydrobiologia 502:169-176. stoyneva m, 2010. [phytoplankton of the wetland pomorie lake, p. 48-60].[article in bulgarian with english abstract]. in: r. radev, g. hiebaum, t. michev and l. profirov (comp.), collection of reports for the integrated management plan for the protexted area pomorie lake bg0000152 and protected area pomorie bg0000620, edition of “zeleni balkani” in the framework of the project “pomorie lake – conservation, restoration and sustainable development” gef 054774. stoyneva mp, 2014. [contribution to the studies of the biodiversity of hydroand aerobiontic prokaryotic and eukaryotic algae in bulgaria].[dr.sc. thesis in bulgarian], sofia university “st kliment ohridski”. stoyneva mp, 2015. allochtonous planctonic algae recorded in bulgaria during the last 25 years and their possible dispersal agents. hydrobiologia 764:53-64. stoyneva m, descy jp, vyverman w, 2007. green algae in lake tanganyika: is morphological variation a response to seasonal changes? hydrobiologia 578:7-16. stoyneva-gärtner m, uzunov b, pavlova v, 2016. [algae in bulgaria as risk factors for human and ecosystem health, p. 15-25].[article in bulgarian with english abstract]. in: proceedings scientific conference “actual problems of safety” 2016, publ. complex of the nvu “vasil levski”. stoyneva mp, traykov it, tosheva ag, uzunov ba, zidarova rp, descy jp, 2015. comparison of ecological state/potential assessment of 19 bulgarian water bodies based on macrophytes and phytoplankton (2011-2012). biotechnol. biotec. eq. 29:33-38. teneva i, basheva d, belkinova d, dimitrova-dyulgerova i, mladenov r, dzhambazov b, 2011. [study of the qualitative and quantitative composition of the blue-green algae (cyanoprokaryota), presence of cyanotoxins and heavy metals in studen kladenets reservoir].[article in bulgarian with english abstract]. plantarum (plovdiv) 41:89-124. teneva i, belkinova d, dimitrova-dyulgerova i, mladenov r, 2009. phytoplankton assemblages and monitoring of cyanotoxins in trakiets reservoir, p. 244-249. in: scientific researches of the union of scientists in bulgaria plovdiv, series b. natural sciences and the humanities, 12, technics, technologies, natural sciences and humanities session, 56 november 2009. teneva i, belkinova d, dimitrova-dyulgerova i, vlaknova m, mladenov r, 2010a. composition and toxic potential of cyanoprokaryota in vacha dam (bulgaria). biotechnol. biotec. eq. 24:26-32. teneva i, mladenov r, belkinova d, dimitrova-dyulgerova i, dzhambazov b, 2010b. phytoplankon community of the drinking after supply reservoir borovitsa (south bulgaria) with an emphasis on cyanotoxins and water quality. centr. eur. j. biol. 5:231-239. teneva i, gecheva g, chesmedjiev s, stoyanov p, mladenov r, belkinova d, 2014. ecological status assessment of skalenski lakes (bulgaria). biotechnol. biotec. eq. 28:82-95. thioulouse j, chessel d, dole s, olivier jm, 1997. ade-4: a multivariate analysis and graphical display software. stat. comp. 7:75-83. traykov i, 2005. [factors influencing the trophic state of the reservoir kardzhali].[phd thesis in bulgarian], sofia university “st kliment ohridski”. tsanev as, belkinova d, 2008. [research on the phytoplankton of ivaylovgrad reservoir (eastern rhodopes mts., bulgaria), p. 485-493].[artcle in bulgarian with english abstract]. in: i.g. velcheva and a.g. tsekov (eds.), proceed. “jubilee scientific ecological conference”, plovdiv. vassilev v, vassilev r, iankov p, kamburova n, uzunov y, pechlivanov l, georgiev b. popgeorgiev g, assyov b, avramov s, tsenova r, kornilev y, 2013. national action plan for conservation of wetlands of high significance of bulgaria 2013-2022. publication of bulgaria biodiversity foundation, sofia: 104 pp. walker hc, 2015. harmful algal blooms in drinking water: removal of cyanobacterial cells and toxins. crc press, boca raton: 145 pp. whitton ba, potts m, 2012. introduction to the cyanobacteria, p. 1-14. in: b.a. whitton (ed.), the ecology of cyanobacteria ii. their diversity in space and time. springer. whitton ba (ed.), 2012. ecology of cyanobacteria ii. their diversity in space and time, springer, dordrecht: 760 pp. who, 1998. guidelines for drinking-water quality, health criteria and other supporting information. 2nd ed. addendum to vol. 2. world health organization, geneva: 253 pp. who, 2003. guidelines for safe recreational water enviroments. algae and cyanobacteria in freshwater. world health organization, geneva: 151 pp. no nco mm er cia l u se on ly layout 1 introduction the field of biodiversity-ecosystem functioning (bef) studies the impact that changes in biological diversity can have on the functioning of ecological systems. for instance, it depicts the consequences of biodiversity loss in terms of basic functions such as nutrient uptake, respiration, primary production and nutrient recycling, among others. in bef research, diversity is quantified generally as the number of species (i.e., richness) or using other richness-based metrics that include species’ abundances (e.g., simpson index, shannon index, evenness). overall, and despite some exceptions, it is well established that most ecosystem functions and their temporal stability increase as the number of species increases (hooper et al., 2005; cardinale et al., 2012; duffy et al., 2017). given the low explanatory power of such richness-based diversity metrics and the absence of a proper mechanistic elucidation, bef research is increasingly adopting a trait-based perspective (flynn et al., 2011; cardinale et al., 2012; krause et al., 2014; gagic et al., 2015). trait variability and the resulting ecological differentiation among species are considered as major determinants of the nature and strength of species interactions and consequently are expected to have a direct strong influence on ecosystem functioning. however, determining and quantifying the traits that are relevant for ecosystem functioning is not straightforward. despite the importance of phytoplankton for global scale processes such as oxygen production and primary production, trait-based bef studies with phytoplankton remain rare. here, i review existing studies linking trait-based diversity to ecosystem functioning in freshwater lentic systems, summarize their major findings and provide some ideas for future development of this underexplored line of research. study selection i collated all the published empirical studies on the relationship between freshwater phytoplankton diversity in lentic systems (lakes, reservoirs, ponds) and any aspect of their functioning. i first collected all previous reviews and meta-analyses on the topic of biodiversity and ecosystem functioning (bef), irrespective of the organism included and checked for references on freshwater phytoplankton (hooper et al., 2005; srivastava and vellend, 2005; balvanera et al., 2006; cardinale et al., 2009; cardinale et al., 2011; cardinale et al., 2013; gross et al., 2014; duffy et al., 2017). this was supplemented with a search of the isi web of science database using the keyword sequence combining (freshwater or lake or pond or reservoir) and (phytoplankton* or alga* or diatom advances in oceanography and limnology, 2017; 8(2): 179-186 review doi: 10.4081/aiol.2017.7207 this work is licensed under a creative commons attribution-noncommercial 4.0 international license (cc by-nc 4.0). biodiversity ecosystem functioning research in freshwater phytoplankton: a comprehensive review of trait-based studies patrick venail1,2 1department f.-a. forel for environmental and aquatic sciences, university of geneva; 2institute for environmental sciences, ise, geneva, switzerland abstract in an effort to reach a clearer mechanistic understanding of the influence of biological diversity on ecosystem functioning, research in the field is increasingly applying a trait-based approach. in this comprehensive review, i searched for and analyzed studies that focused on the relationship between biodiversity and ecosystem functioning (bef) using a trait-based approach in freshwater phytoplankton from lentic systems (lakes, ponds, reservoirs). i found that this type of studies is very rare and included a plethora of traits, diversity metrics, statistical analyses and study locations that contributed to the high variability in the results they obtained. overall, trait-based diversity is not a very good predictor of ecosystem functioning in freshwater lentic ecosystems. null relationships between trait-based diversity and ecosystem functioning in freshwater lentic systems were the more frequent outcome. when significant, the amount of variation in ecosystem functioning explained by trait-based diversity was small. still, trait-based research remains a promising approach to increase the mechanistic understanding of bef relationships. for this purpose, studies directly testing the underlying mechanistic rationale, exploring diversity effects on the temporal stability of ecosystem functions, including multiple functions at a time, focusing more in cell size and shape and confirming the relative importance of individual trait variation for ecosystem functioning are needed. key words: biodiversity; freshwater; functioning; phytoplankton; traits. received: november 2017. accepted: december 2017. no nco mm er cia l u se on ly p. venail180 or cyano*) and (trait* or function*) and (diversity or richness) and (community or ecosystem) and (function* or product* or biomass or biovolume or resource use). in this review, i only included studies that statistically analyzed the link between any trait-based metric of diversity and ecosystem functioning using freshwater phytoplankton. i excluded studies in lotic systems (i.e., rivers and streams) and studies in which phytoplanktonic organisms were classified into functional groups, such as major algal groups or reynold’s classification because they did not include a clear trait-based diversity measure (schmidtke et al., 2010; behl et al., 2011; borics et al., 2012; fernandez et al., 2014; abonyi et al., 2017). studies i found only six studies that explored the relationship between trait-based diversity and ecosystem functioning using freshwater phytoplankton from lentic systems (tab. 1). these studies are recent, with the oldest published only seven years ago (vogt et al., 2010), revealing that, on average, less than one study per year has been published in this topic. the six studies can be separated into two categories: field and laboratory studies. the former includes four studies in which both trait-based diversity and ecosystem functioning data were obtained from natural conditions (vogt et al., 2010; pälffy et al., 2013; santos et al., 2014; fontana et al., 2017). in the other two studies, diversity was directly manipulated under highly controlled conditions in the laboratory (shurin et al., 2014; steudel et al., 2016). the main conclusions of these studies regarding the link between trait-based diversity and functioning are quite variable (tab. 1). below, i explore those differences more in detail. organisms and freshwater systems the six studies reported a wide variety of organisms from all major phytoplankton groups including mostly chlorophyta (green algae), chrysophyta (golden algae), bacillariophyta (diatoms) and cyanophyta (cyanobacteria). phytoplankton from other major groups such as glaucophyta and heterokontophyta were less frequent. the richness of taxa within each study was rather variable. the laboratory studies included 16 (shurin et al., 2014) and 64 (steudel et al., 2016, only chlorophytes) taxa respectively, whereas field studies reported 212 (vogt et al., 2010) and 412 species (santos et al., 2014). the other two field studies did not report the number of species analyzed. field data were collected in a variety of freshwater systems around the world: vogt et al. (2010) included data from 65 lakes in canada, pälffy et al. (2013) from one single lake in hungary, santos et al. (2014) included 19 reservoirs from brazil, whereas fontana et al. (2017) reported data from 28 lakes in total, 2 from switzerland and 26 from the danube delta in romania. traits a total of 33 traits were included in the six studies about the effect of trait-based phytoplankton diversity on ecosystem functioning in freshwater lentic systems (tab. 2). the authors selected such traits based on their supposed ecological relevance for competitive interactions, reproduction, predator avoidance, resource acquisition and/or bioenergy production. most of these traits were measured at the species level, meaning that they represent an average value obtained by measuring and recording traits in some representative individuals or populations from each species. different to all other studies, fontana et al. (2017) recorded individual level data for seven traits, meaning that species’ identification was not required prior to trait measurement and the reported trait values may reveal both intra and inter-specific variability in the phytoplankton community. traits reported belong to three different categories: demographic, morphological or physiological. demographic tab. 1. list of trait-based bef studies in freshwater lentic systems, type of study and their main conclusions regarding trait-based diversity effects on ecosystem functioning. reference type of study main conclusion vogt et al., 2010 field trait-based diversity was positively associated with total community biovolume pälffy et al., 2013 field significant negative correlations between total biomass, functional group diversity and functional group evenness santos et al., 2014 field a positive relation between productivity and diversity, except for functional evenness for which the relation was negative fontana et al., 2017 field trait evenness exhibited a robust negative relationship with biomass shurin et al., 2014 laboratory biomass yield exceeded the component monocultures in polycultures consisting of species with highly divergent traits steudel et al., 2016 laboratory functional diversity was positively correlated with biomass overyield bef, biodiversity and ecosystem functioning. no nco mm er cia l u se on ly biodiversity ecosystem functioning research in freshwater phytoplankton: a comprehensive review of trait-based studies 181 (7 traits), also named life history traits, are all continuous and include population growth parameters such as r or k (measured by using chlorophyll a as a proxy for growth; shurin et al., 2014). in another study, vogt et al. (2010) included five demographic traits reported as “response traits” because they are based on how different environmental parameters influence population growth. these include: optimal growth conditions regarding total nitrogen, total phosphorous, ph, dissolved organic carbon and dissolved co2. morphological (10 traits), include continuous, categorical or binomial traits. in this category, we find cell size, also reported in some studies as cell volume, greatest axial linear dimension (gald), maximal linear dimension (mld) or maximum length. this is the more frequently used trait, included in five different studies (tab. 2). while cell size is normally reported as a continuous variable, palffy et al. (2013) reported it as a categorical variable with three size classes. other morphological traits that are reported in multiple independent studies are growth form (referring to either colonial or single cell organisms), presence of gas vacuoles (referring to buoyancy control capabilities), and the presence of flagella, that relates to motility. the other six traits in the morphological category were only reported in one independent study and can be either binomial or continuous (tab. 2). physiological (16 traits) represent the largest array of features among the three categories and can be either continuous, categorical or binomial and each was reported in only one study. these traits relate to minimum resource requirements (light, nitrogen, phosphorous, silica), cellular chemical content, biochemistry or stoichiometry (lipids, fatty acid, carbon, nitrogen, phosphorous), resource actab. 2. list of traits included in bef studies with freshwater phytoplankton. demographic er type reference exponential growth rate, r* bp, ci continous shurin et al., 2014 asymptotic density, k* bp, ci continous “ total nitrogen optimal concentration continous “ total phosphorous optimal concentration continous “ ph optimal continous “ dissolved organic carbon optimal continous “ dissolved co2 optimal continous “ morphological cell volume/size/gald/mld/max length pa, ra, re, bp, ci continous/categorical all but steudel et al., 2016 growth form/body form/complexity pa, ra, re categorical pälffy et al., 2013; santos et al., 2014 surface to volume ratio, s/v continous pälffy et al., 2013 presence of aerotopes/gaz vacuoles/buoyancy pa, ra binomial/categorical pälffy et al., 2013/santos et al., 2014 presence of flagella/motility pa, ra binomial “ presence of mucilage pa, ra binomial santos et al., 2014 presence of siliceous exoskeletal structures pa, ra binomial “ presence of heterocysts pa, ra binomial “ frontal shape of particle ra continous fontana et al., 2017° cell rugosity/internal structure/gas vesicle/thylacoids pa, ra continous “ physiological cellular lipid concentration bp, ci continous shurin et al., 2014 cellular c:n ratio bp, ci continous “ cellular c:p ratio bp, ci continous “ minimum light requirement, l* bp, ci continous “ minimum nitrogen requirement, n* bp, ci continous “ minimum phosphorous requirement, p* bp, ci continous “ fatty acid composition ci continous steudel et al., 2016 photosynthetic pigment composition categorical pälffy et al., 2013 fluorescence chlorophyl a ra continous fontana et al., 2017° fluorescence phycoerythrin ra continous “ fluorescence accesory pigments ra continous “ eveness in the distribution of pigments within cell ra continous “ hability to fix nitrogen binomial pälffy et al., 2013 phagotrophic potential binomial “ motility/buoyancy categorical “ presence of toxins pa binomial santos et al., 2014 bef, biodiversity and ecosystem functioning; er, ecological relevance as explicitly claimed by the authors; ra, resource acquisition; re, reproduction; pa, predator avoidance; bp, bioenergy production; ci, competitive interactions; *growth using fluorescence (chl a); °individual level traits. no nco mm er cia l u se on ly p. venail182 quisition (pigments, phagotrophy, nitrogen fixation) and toxin production. trait-based diversity variables and metrics the six studies reviewed here include 29 trait-based variables that were associated to ecosystem functioning afterwards (tab. 3). these variables can be classified in three categories: functional group based variables, traitbased diversity metrics and trait-based non-diversity metrics. the functional group category includes well-known diversity metrics that are traditionally used to quantify species-level diversity such as richness, shannon index, simpson index and evenness (the three latter incorporate information on species’ abundances). for this, species are first classified into functional groups such as those proposed by kruk (kruk et al., 2010) or in major algae groups (e.g., chlorophytes or cyanophytes). then, the different metrics were calculated based on group richness’ information. kruk’s classification is based on morphological aspects and consequently corresponds to actual traitbased quantification of diversity; classifications based on major algae groups or reynold’s groups are not traitbased only (reynolds et al., 2002) and as such should not be considered as formal traits-based diversity metrics. the other two categories of trait-based metrics require collecting trait information on species (but see fontana et al., 2017 for individual level trait metrics). species’ traits are aggregated according to the taxa present in the natural community or artificial assemblage. this aggregation may include averaging, calculating distances or variation among species and other more sophisticated aggregation methods. the trait-based diversity metrics are the more common and diverse in the phytoplankton bef literature, as i recorded up to 20 different metrics in five studies (tab. 3). some metrics incorporate information of one single trait (11 in total) at a time while others include up to six (vogt et al., 2010), seven (fontana et al., 2017), eight (santos et al., 2014) or nine traits simultaneously (shurin et al., 2014). this type of metrics can also be weighted by species’ abundances. none of the twenty different diversity metrics based on traits were used in more than one study, revealing a large variability in the methodology of trait-based bef studies. up to six different metrics were used in one single study (vogt et al., 2010). tab. 3. list of trait-based variables associated to ecosystem functioning in freshwater phytoplankton. functional groups reference functional group richness (kruk’s groups, taxonomic) santos et al., 2014 functional group diversity (shannon hf) pälffy et al., 2013 functional group diversity (evenness jf) “ functional group diversity (simpson) santos et al., 2014 trait-based diversity metric difference in pca vector (on 9 traits) shurin et al., 2014 functional dispersion, fdis (on 9 traits) “ difference in c:n ratios between 2 species “ difference in cell volume between 2 species “ difference in minimun light requirement l* between 2 species “ fatty acid composition similarity, ftd steudel et al., 2016 fatty acid composition similarity, fd “ variance of species in total nitrogen optima, tv vogt et al., 2010 variance of species in total phosphorous optima, tv “ variance of species in ph optima, tv” variance of species in dissolved organic carbon optima, tv “ variance of species in co2 optima, tv” sum branch length dendrogram (on 6 traits) “ functional richness based on distances, fr (on 8 traits) santos et al., 2014 functional evenness based on distances, feve (on 8 traits) “ functional divergence based on distances, mfd (on 8 traits) “ functional divergence weigthed by density, mfddens (on 8 traits) “ trait diversity richness, top (on 7 traits) fontana et al., 2017 trait diversity evenness, ted (on 7 traits) “ trait diversity divergence, fdis (on 7 traits) “ other trait based gradients (not variation) average pca vector (on 9 traits) shurin et al., 2014 average cell volume of 2 species “ average c:p ratios of 2 species “ average c:n ratios of 2 species “ average minimum phosphorous requirement p* of 2 species “ no nco mm er cia l u se on ly biodiversity ecosystem functioning research in freshwater phytoplankton: a comprehensive review of trait-based studies 183 finally, the third category includes trait-based metrics that do not represent diversity per se because they are just average traits among species and do not include information in the variation of a trait (shurin et al., 2014). some studies combine multiple of these metrics together in one single statistical analysis to determine the combinations of metrics describing better the variation in ecosystem functioning among communities (santos et al., 2014; fontana et al., 2017). ecosystem functioning five different variables were documented as measures of ecosystem functioning and can be separated in two categories: the biomass related and the non-biomass related. the two controlled laboratory studies (shurin et al., 2014; steudel et al., 2016) focused on biomass related ecosystem functioning variables. for this, they started by measuring the biomass of both polycultures and monocultures using optical density as a proxy. then, two log-ratios were calculated: one between the biomass of the polyculture to the average of constitutive monocultures (i.e., net biodiversity effect, nbe) and another between the biomass of the polyculture to the more productive monocultures (i.e., overyielding, oy). this method can only be applied to controlled laboratory experiments because it requires monoculture’s biomass estimations. the third biomass related variable, used in two field studies was total community biomass (vogt et al., 2010; fontana et al., 2017). the two nonbiomass related ecosystem functions included in the other field studies are chlorophyll a concentration (pälffy et al., 2013; santos et al., 2014) and a proxy for resource use efficiency calculated as the ratio between total biomass and available total phosphorous (fontana et al., 2017). biodiversity and ecosystem functioning relationships a key step in every bef study is to relate biodiversity (the explanatory variable) to ecosystem functioning (the response variable). in controlled laboratory studies, this link infers causality because all the observed variations in ecosystem functioning result from changes in either the diversity and/or the composition of the species assemblages being tested. in field studies, given the possibility of abiotic and biotic changes among sites or dates, the link between diversity and ecosystem functioning is just correlational. a plethora of statistical methods have been used to relate diversity to ecosystem functional as causality effects or correlational links, including correlations, linear regressions and linear mixed effect models. the latter allows combining multiple diversity metrics in one single statistical model (steudel et al., 2016; fontana et al., 2017). a total of 190 relationships between trait-based diversity and ecosystem functioning have been established so far for freshwater phytoplankton (tab. 4). i classified them into either positive, null or negative based on the statistical analyses directly reported by the authors. half of them showed no influence of trait-based diversity on ecosystem functioning, meaning that variations in functioning are independent from variations in trait diversity among freshwater phytoplankton. this higher prevalence of null relationships, compared to the significant ones, was consistent in both field and laboratory studies with 42% and 53% of total bef relationships being null respectively. positive bef relationships, meaning that ecosystem functions considered increase as phytoplankton trait diversity increases, were present in nearly 40% of the experiments. the authors suggest some potential mechanisms to explain this positive effect of diversity. a larger functional trait-based diversity may embrace a wider range of optimal growth responses in face of changing abiotic lake conditions, leading to higher biomass production (vogt et al., 2010). also, species with different traits would be more complementary in resource use or compete less strongly against each other (santos et al., 2014; shurin et al., 2014). negative relationships were present in 11.5% of cases, which means a decreasing ecosystem functioning as trait diversity increased. as an explanation for this pattern, authors suggest that under some conditab. 4. summary table of the effect of trait-based metrics on ecosystem functioning. type of study reference sign of trait based diversity effects on function positive null negative field vogt et al., 2010 19 11 0 pälffy et al., 2013 0 0 4 santos et al., 2014 2 1 1 fontana et al., 2017 6 19 11 laboratory shurin et al., 2014 8 2 6 steudel et al., 2016 40 60 0 total 75 93 22 percentage 39.5% 49% 11.5% no nco mm er cia l u se on ly p. venail184 tions one single productive taxa with particular traits might dominate, leading to low functional diversity coupled to high biomass (pälffy et al., 2013;santos et al., 2014). unfortunately, none of the studies reviewed here explicitly tested the suggested mechanisms and remained purely conjectural. some authors also suggest that other forces (such as resource scarcity) might simultaneously influence both trait diversity and ecosystem functioning, resulting in a negative or positive pattern that is not mediated directly by diversity (fontana et al., 2017). sources of variation in bef relationships the reviewed papers allowed determining a series of factors influencing the relationship between trait-based diversity and ecosystem functioning in experiments with freshwater phytoplankton. vogt et al. (2010) found big differences in sign and strength of the bef relationship depending on: the trait, the number of traits, the metric of diversity and the organisms included in the analysis. for instance, whereas functional diversity had a positive effect on total biovolume in the benthic algae, this effect was not present in the planktonic compartment. the inverse happened for functional diversity based on one single trait (i.e., co2 optimal). fontana et al. (2017) showed that the relationship between trait-based diversity and ecosystem functioning may also vary among locations. for instance, a combination of three trait-based diversity metrics describes very well variations in biomass in lake greifensee but this same set of metrics describes less well total biomass in lake zurich or the danube delta. steudel et al. (2016) showed that the effects of trait-based diversity also depend on the number of interacting species, with a tendency for higher influence of trait-based diversity as the number of species increases from two to sixteen. best trait-based descriptors of ecosystem functioning one purpose of bef studies is to determine which trait (or set of traits) and which metric (or set of metrics) describes better variations in ecosystem functioning. in case of significant effects (p<0.05), either positive or negative, one wants to know which trait is the best predictor using coefficients of variance (r2 values) and to compare models using for instance the akaike information criterium (aic). this would reveal which trait matters the most for ecosystem functioning in freshwater lentic ecosystems. one may as well want to know which traits do not influence ecosystem functioning. i collected data on the percentage of variation in ecosystem functioning explained by trait-based diversity and found that this percentage ranged from 1.8% to 90%, with an average of 34.7% (n = 51). the model that explained the most (90%) of the variation in ecosystem functioning included one single metric of diversity (i.e., trait diversity evenness, ted), which was based on individual level trait variation and included seven different traits (fontana et al., 2017). in studies focused on species level trait variation (5 out of 6 studies), the best single metric describing ecosystem functioning included information from six traits and explained 54% of variation in total biomass production in benthonic diatom communities (vogt et al., 2010). the same study showed that the capacity to predict biomass production depended on the number of traits considered. this is, including two or three traits, rather than five or just one, increased the predictive power. this result suggests that different traits may incorporate different information but also that some traits might be functionally redundant and their use may not lead to a better description of biomass variation among communities. gene-based diversity as a proxy for trait-based diversity one study explored the capacity of gene-based diversity to predict freshwater phytoplankton functioning (steudel et al., 2016). the underlying rationale is that gene differentiation among species (i.e., phylogenetic divergence) may relate to trait differentiation (assuming phylogenetic signal or phylogenetic niche conservatism; blomberg and garland 2002; losos 2008; wiens et al., 2010). overall, the results show that the explanatory power is higher for trait-based diversity metrics than for gen-based diversity metrics. moreover, at high richness level (16 species) both types of metrics have contrasting effects on biomass production. trait-based metrics had a positive effect on biomass whereas gene-based metrics had a negative effect on biomass. this suggests that genebased metrics should not be considered as proxies of traitbased ones in freshwater communities. ideas for future development in the biodiversity-ecosystem functioning context, stability refers to the capacity of an ecological system to perform ecosystem functions despite variations (e.g., perturbations) in the abiotic or biotic conditions over time. overall, diversity is expected to have a strong positive impact on ecosystem functioning stability given the capacity of a diverse set of organisms to cope with the different environmental conditions (hooper et al., 2005; tilman et al., 2006). freshwater lentic ecosystems are steadily exposed to changing environmental conditions such as temperature, light intensity, resource input, etc. no study has no nco mm er cia l u se on ly biodiversity ecosystem functioning research in freshwater phytoplankton: a comprehensive review of trait-based studies 185 explored the influence of trait-based diversity on the temporal stability of ecosystem functioning. trait-based bef studies in freshwater phytoplankton focused on single functions. however, ecological systems often perform multiple functions at a time and some of these functions are expected to be directly related, such as resource uptake and biomass production. it has been suggested that diversity effects on ecosystem functioning might be stronger when multiple functions are considered simultaneously (byrnes et al., 2014; lefcheck et al., 2015). such a multifunctional approach using trait-based diversity in freshwater lentic systems is missing. cell size and shape are considered as key traits for phytoplankton, directly influencing resource acquisition, reproduction, predator avoidance and species interactions (litchman and klausmeier, 2008; finkel et al., 2010). whereas some trait-based bef studies reported here included phytoplankton cell size and/or shape in the estimation of trait-based diversity, only one study (shurin et al., 2010) explored directly the effect of these traits on ecosystem functioning. more studies manipulating phytoplankton cell size and/or shape are required to provide more consistent conclusions on this topic. individual trait information may increase the descriptive power of ecosystem functioning (fontana et al., 2017). more studies incorporating and comparing the explanatory capacity of individual vs. species trait-based metrics of diversity on ecosystem functioning are also required to determine the generality of this effect. finally, motivation to include trait-based information into bef research supposes that trait variability among species reflects their ecological differentiation and thus determine the nature and strength of species interactions that ultimately influence ecosystem functioning. trait-based studies reported to date did not explicitly test this rationale and the proposed mechanistic interpretations remain purely conjectural. studies directly testing the underlying mechanisms such as the traits involved in the prevalence of competition or facilitation should be further developed. in a nutshell this first comprehensive review of studies linking traitbased freshwater phytoplankton in lentic systems to ecosystem functioning revealed the scarce research conducted in this topic, with only six published studies over the last seven years. two of these studies were conducted under controlled laboratory conditions and the other four studies reported field collected data, in which other external variables might be influencing biodiversity-ecosystem functioning relationships. a total of 33 traits and 29 diversity metrics have been reported. traits are either demographic, morphological or physiological. some diversity metrics are based in one single trait but the majority include several traits simultaneously. no empirical evidence suggests that variation in one specific trait or group of traits improves the predictability of ecosystem functioning in freshwater lentic systems. similarly, including multiple traits simultaneously or including multiple trait-based diversity metrics together does not necessarily make bef relationships stronger. nevertheless, a plethora of traits, diversity metrics, statistical analyses and study locations contributed to the high variability in the results obtained. null relationship between trait-based diversity and ecosystem functioning in freshwater lentic systems was the more frequent outcome, accounting for nearly half of the experiments. when statistically significant, positive effects of trait-based diversity on ecosystem functioning were nearly four times more common than negative ones. in these studies, the amount of variation in ecosystem functioning explained by traitbased diversity was variable but rather small. overall, this means trait-based diversity is often not a very good predictor of ecosystem functioning in freshwater lentic ecosystems. the capacity to improve our mechanistic understanding of biodiversity-ecosystem functioning relationships in freshwater lentic ecosystems using trait information has not been fully exploited. studies directly testing the underlying mechanistic rationale are required. other ideas for further development in this field include studying diversity effects on the temporal stability of ecosystem functions, exploring multiple functions at a time (multifunctionality), focusing exclusively in cell size and shape as master traits and confirming the relative importance of individual trait variation for ecosystem functioning. references abonyi a, horváth z, ptacnik r, 2017. functional richness outperforms taxonomic richness in predicting ecosystem functioning in natural phytoplankton communities. freshwater biol. doi: 10.1111/fwb.13051. [epub ahead of print]. balvanera p, pfisterer ab, buchmann n, he js, nakashizuka t, raffaelli d, schmid b, 2006. quantifying the evidence for biodiversity effects on ecosystem functioning and services. ecol. lett. 9:1146-1156. behl s, donval a, stibor h, 2011. the relative importance of species diversity and functional group diversity on carbon uptake in phytoplankton communities. limnol. oceanogr. 56:683-694. blomberg sp, garland jrt, 2002. tempo and mode in evolution: phylogenetic inertia, adaptation and comparative methods. j. evol. biol. 15:899-910. borics g, tóthmérész b, lukács ba, várbíró g, 2012. functional groups of phytoplankton shaping diversity of shallow lake ecosystems. hydrobiologia 698:251-262. byrnes jek, gamfeldt l, isbell f, lefcheck js, griffin jn, hector a, cardinale bj, hooper du, dee le, duffy e, 2014. investigating the relationship between biodiversity and ecosystem multifunctionality: challenges and solutions. methods ecol. evol. 5:111-124. no nco mm er cia l u se on ly p. venail186 cardinale bj, duffy je, gonzalez a, hooper du, perrings c, venail p, narwani a, mace gm, tilman d, wardle da, kinzig ap, daily gc, loreau m, grace jb, larigauderie a, srivastava ds, naeem s, 2012. biodiversity loss and its impact on humanity. nature 486:59-67. cardinale bj, gross k, fritschie k, flombaum p, fox jw, rixen c, van ruijven j, reich pb, scherer-lorenzen m, wilsey bj, 2013. biodiversity simultaneously enhances the production and stability of community biomass, but the effects are independent. ecology 94:1697-1707. cardinale bj, matulich kl, hooper du, byrnes je, duffy e, gamfeldt l, balvanera p, o’connor mi, gonzalez a, 2011. the functional role of producer diversity in ecosystems. am. j. botany 98:572-592. cardinale bj, srivastava ds, duffy je, wright jp, downing al, samkaran m, jouseau c, cadotte mw, carroll it, weis jj, hector a, loreau m, 2009. effects of biodiversity on the functioning of ecosystems. a summary of 164 experimental manipulations of species richness. ecology 90:854. duffy je, godwin cm, cardinale bj, 2017. biodiversity effects in the wild are common and as strong as key drivers of productivity. nature 549:261-264. fernandez c, cáceres ej, parodi er, 2014. phytoplankton development in a highly eutrophic man-made lake from the pampa plain of argentica a functional approach. int. j. environ. res. 8:1-14. finkel zv, beardall j, flynn kj, quigg a, rees tav, raven ja, 2010. phytoplankton in a changing world: cell size and elemental stoichiometry. j. plankt. res. 32:119-137. fontana s, thomas mk, moldoveanu m, spaak p, pomati f, 2017. individual-level trait diversity predicts phytoplankton community properties better than species richness or evenness. isme j. doi: 10.1038/ismej.2017.160. [epub ahead of print] flynn dfb, mirotchnick n, jain m, palmer mi, naeem s, 2011. functional and phylogenetic diversity as predictors of biodiversity ecosystem-function relationships. ecology 92: 1573-1581. gagic v, bartomeus i, jonssin t, taylor a, winqvist c, fischer c, slade em, steffan-dewenter i, emmerson m, potts sg, tscharntke t, weisser w, bommarco r, 2015. functional identity and diversity of animals predict ecosystem functioning better than species-based indices. proc. r. soc. b 282: 20142620. gross k, cardinale bj, fox jw, gonzalez a, loreau m, polley hw, reich pb, van ruijven j, 2014. species richness and the temporal stability of biomass production: a new analysis of recent biodiversity experiments. am. nat. 183:1-12. hooper du, chapin fs, ewel jj, hector a, inchausti p, lavorel s, lawton jh, lodge dm, loreau m, naeem s, schmid b, setalah, symstad aj, vandermeer, wardle da, 2005 effects of biodiversity on ecosystem functioning: a consensus of current knowledge. ecol. monog. 75:3-35. krause s, le roux x, niklaus pa, van bodegom pm, lennon jt, bertilsson s, grossart h-p, philippot l, bodelier pl, 2014. trait-based approaches for understanding microbial biodiversity and ecosystem functioning. front. microbiol. 5:251. kruk c, huszar vlm, peeters ethm, bonilla s, costa l, lürling m, reynolds c, scheffer m, 2010. a morphological classification capturing functional variation in phytoplankton. freshwater biol. 55:614-627. lefcheck js, byrnes jek, isbell f, gamfeldt l, griffin jn, eisenhauer n, hensel mjs, hector a, cardinale bj, duffy e, 2015. biodiversity enhances ecosystem multifunctionality across trophic levels and habitats. nature comm. 6:6936. litchman e, klausmeier ca, 2008. trait-based community ecology of phytoplankton. annu. rev. ecol. evol. syst. 39:615-639. losos jb, 2008. phylogeneitc niche conservatism, phylogenetic signal and the relationship between phylogenetic relatedness and ecological similarity among species. ecol. lett. 11:9951007. pälffy k, présing m, vörös l, 2013. diversity patterns of traitbased phytoplankton functional groups in two basins of a large, shallow lake (lake balaton, hungary) with different trophic state. aquat. ecol. 47:195-210. reynolds cs, huszar v, kruk c, naselli-flores l, melo s, 2002. towards a functional classification of the freshwater phytoplankton. j. plankt. res. 24:417-428. santos amc, carneiro fm, cianciaruso mv, 2014. predicting productivity in tropical reservoirs: the roles of phytoplankton taxonomic and functional diversity. ecol. ind. 48:428-435. schmidtke a, gaedke u, weithoff g, 2010. a mechanistic basis for underyielding in phytoplankton communities. ecology 91:212-221. shurin jb, mandal s, abbott rl, 2014. trait diversity enhances yield in algal biofuel assemblages. j. appl. ecol. 51:603-611. srivastava ds, vellend m, 2005. biodiversity-ecosystem function research: is it relevant to conservation. annu. rev. ecol. evol. syst. 36:267-294. steudel b, hallmann c, loremz m, abrahamczyk s, prink k, herrfurth c, feussner i, martini jwr, kessler m, 2016. contrasting biodiversity-ecosystem functioning relationships in phylogenetic and functional diversity. new phytologist 212:409-420. tilman d, reich pb, knops jm, 2006. biodiversity and ecosystem stability in a decade-long grassland experiment. nature 441:629-632. vogt rj, beisner be, prairie yt, 2010. functional diversity is positively associated with biomass for lake diatoms. freshwater biol. 55:1636-1646. wiens jj, ackerly dd, allenn ap, anacker bl, buckley lb, cornell hv, damschen ei, davies tj, grytnes j-a, harrison sp, hawkins ba, holt rd, mccain cm, stephens pr, 2010. niche conservatism as an emerging principle in ecology and conservation biology. ecol. lett. 13:1310-1324. no nco mm er cia l u se on ly layout 1 introduction the application of the multiproxy approach in paleolimnological studies is becoming increasingly widespread as it allows to integrate the responses of lakes and their catchment and to disentangle the effects of combined human and climate impacts (batterbee, 2000; bennion et al., 2015). the combination of biological and geochemical proxies allows tracking different aspects of specific impacts (bennion et al., 2015; perga et al., 2015), and overtaking the limits of the different studied proxies, such as cladocera, diatoms, subfossil pigments and lithogenic elements (rosen et al., 2010). cladocera represent key players of the lake food web as they act between top-down regulators (fish and invertebrate predators) and bottom-up factors (nutrients and phytoplankton; eggermont and martens, 2011). in particular, the study of cladocera remains allows to reconstruct changes in lake trophic status (jeppesen et al., 2001), water temperature (nevalainen, 2012; szeroczyńska, 2006) and ph (jeziorski et al., 2008), water level (korhola et al., 2005), macrophyte distribution (davidson et al., 2007), and food web (finney et al., 2000; jeppesen et al., 2001). the studies by manca et al. (2007) and korosi et al. (2010) on changes in cladocera body and appendages length over longer periods highlighted that size varies in relation to predators pressures, water temperature, nutrient, ph and ca. these analyses resulted to be particularly useful when long-term data on fish and invertebrate predators are not available, as well as to support information provided by other proxies. geochemical analyses of lake sediments have been advances in oceanography and limnology, 2016; 7(2): 220-234 article doi: 10.4081/aiol.2016.6399 this work is licensed under a creative commons attribution-noncommercial 4.0 international license (cc by-nc 4.0). combining sediment cladocera remains and geochemistry to reveal the role of a large catchment in driving changes in a small subalpine lake (lake ledro, n-italy) manuela milan,1* richard bindler,1 monica tolotti2 1department of ecology and environmental sciences, umeå university, linnaeus väg 6, 90187 umeå, sweden 2sustainable ecosystems and bioresources, research and innovation centre, fondazione e. mach, via e. mach 1, 38010 san michele all'adige (tn), italy *corresponding author: milan.manuela@gmail.com abstract sediment cladocera remains and geochemistry were analyzed at lake ledro, a small subalpine lake with a large catchment area located in northern italy. the aim of the study was to investigate human, climate and hydrological impacts on the cladocera community and on the geochemical components during the last few centuries. a sediment core was collected from the deepest point of lake ledro and radiometrically dated. cladocera remains were analyzed to track the trophic lake evolution. the core bottom section revealed the dominance of bosminidae in concomitance with nutrient pulses entering into the lake during major flood events. the abundance of species preferring cold water temperatures confirmed the deposition of this core section during the little ice age. the flood event occurred in the first half of the 19th century produced a drastic increase in littoral species, due to the development of new habitats. the decrease in cladocera densities during the following lake stage was followed by a rapid increase in planktonic species during the nutrient enrichment after the 1960s. statistical analyses revealed a clear response of cladocera community to climate variability during oligotrophic periods, while no relation to temperature changes was recorded during high nutrient levels. a preliminary study on bosminidae and daphnidae body size and appendages length was carried out to reconstruct major changes in the lake food web. only bosmina spp. revealed clear body size changes: minor shifts were recorded before the 1930s in relation to the low nutrient concentrations, while the major changes occurred during the 1980s were interpreted as related to the appearance of cladocera invertebrate predators. geochemical components were studied using x-ray fluorescence spectroscopy (xrf) analysis in order to recognize the impact of the large catchment area and from the lake-level regulations on the lake hydrology. moreover the si:al ratios profile confirmed the increase in lake productivity after the 1960s. although both cladocera and geochemical analysis indicate major changes since the 1960s, they also revealed diverse responses to common external and local forcing, thus confirming the value of a multi-proxy approach for disentangling the lake responses to different environmental stressors. moreover, it outlined the importance of larger catchment areas on small lakes as they are to a larger extent influenced by the modifications occurring in the drainage basin. key words: cladocera; geochemistry; wavelenght-dispersive x-ray fluorescence spectroscopy; paleolimnology; hydrological impact; bosmina morphology. received: november 2016. accepted: december 2016. no n c om me rci al us e o nly sediment cladocera and geochemistry of lake ledro 221 successfully used to reveal responses of lake ecosystems to catchment change (battarbee, 2000). in fact, soil stabilization and erosion are strongly influenced by land-use (garcía-ruiz et al. 2010), while physical and chemical weathering of the catchment area are affected by climatic variability (vannière et al., 2013). the analysis of different geochemical components can provide information on lake productivity (e.g., si:al as a proxy of biogenic silica), redox conditions (i.e., fe:mn ratios), atmospheric pollution (e.g., pb concentration), land use (lithogenic elements), climate change and weathering rates (k:al ratios; martín-puertas et al., 2011). several paleolimnological studies were conducted on subalpine lakes located on the northern slope of the alps (alefs and müller, 1999, berthon et al., 2013; wessels et al., 1999), while only a few sediment records were studied on the southern side (marchetto et al., 2004; guilizzoni et al., 2006; milan et al., 2015). as already anticipated by battarbee (2000), the recent work on the largest italian lake, lake garda (milan et al., 2015), highlighted the necessity to expand the study to other lakes of the italian subalpine region in order to understand both the individualistic response to common external forcing, and the importance of local factors in driving the lake dynamics at secular scale. lake ledro appeared to be the ideal site to begin this expansion as it is located very close to lake garda and its large drainage basin has a great influence on the lake hydrology. several physical analyses were recently conducted on the sediments of this lake in order to understand the impact of flood events on the lake (magny et al., 2012; vannière et al., 2013; simonneau et al., 2013b). however, the response of the biological communities to long term changes is still unknown as sediments of lake ledro were never analyzed for biological proxies. the aim of this work was to reconstruct the long-term influence of human, climate and hydrological impacts on biological and geochemical components of lake ledro during the last few centuries. the multiproxy approach was applied in order to identify and disentangle the effects of the different impacts on the lake-catchment systems. this appears to be of particularly importance for this lake as its large drainage basin is related to major source of disturbance. cladocera remains were identified in the sediment samples and related with independent environmental and climatic variables in order to reconstruct the influence on the lake biological community exerted by major climate changes as occurred for example during the little ice age, by the post-war nutrient enrichment and by major flood events. a preliminary study on the body size and appendages length was carried out in order to overtake the lack of long-term predator information. finally, sediment samples were analyzed by x-ray fluorescence spectroscopy (xrf) for the reconstruction of the geochemical changes. the different elements and ratios have been compared in order to understand magnitude and impacts on the lake dynamics of major flood events and water-level regulations. methods study site lake ledro is a small glacial lake (area: 3.7 km2, vol: 0.08 km3, zmax: 49 m) with a catchment area of ca. 111 km2, ranging from 2254 m asl down to 652 m asl (fig. 1). several stages of high flood frequency during the holocene were outlined by previous studies on lake ledro sediments and interpreted in relation to combined effects of the torrential regime of the two temporary tributaries, massangla and pur rivers, the steepness of their valleys and the high ratio between catchment and lake area (30:1 ratio, vannière et al., 2013; simonneau et al., 2013b). lake ledro is situated very close to the northern extremity of lake garda (fig. 1), the largest italian lake, to which it is connected through an underwater pipe located at 25 m depth in lake ledro. water is forced through a pump-storage power plant built on the river ponale down to lake garda, and then the water is pumped back up. the lake-level of lake ledro has been regulated for hydroelectricity production since ad 1929 (vannière et al., 2013). sediment coring and chronology a gravity corer (uwitec, austria) was used to collect a short core (83 cm) from the deepest point of lake ledro (45°52’44”n, 10°45’10”e) in december 2011. the core was vertically extruded and sliced in the laboratory at 0.5 cm intervals from 0 to 30 cm, and at 1 cm intervals from 31 cm down to the core bottom. sediment aspect and texture were annotated during the slicing. the core chronology was established applying the crs dating model (appleby, 2001) to direct gamma assay radiometric analyses of210pb, 226ra, 137cs and 241am, which were conducted at ensis ltd-university college london, uk. geochemistry and subfossil pigments wet density (wd), water content (h2o, measured from dry weight) and total organic matter (om, measured as loss on ignition at 550°c) were determined for all the subsamples as described in milan et al. (2015). about 0.2 g of dried and homogenized sediment were used to measure major and trace elements using a bruker s8 tiger wd-xrf analyzer equipped with an rh anticathode xray tube (further detail in rydberg, 2014). the analysis showed accuracy within 10% and analytical precision within 5% for the majority of the elements. concentrations of key groups (i.e., major elements, redox, lithologno n c om me rci al us e o nly 222 m. milan et al. ical and trace elements) and ratios between specific elements are discussed in this study. in particular, the mn:fe ratio was used to infer changes in the sediment redox conditions, while the si:al ratio was used to infer changes in biogenic silica. the pb enrichment factor (pbef) was calculated in order to outline the atmospheric pollution in this region by using titanium (ti) as the reference element. the value of pb:ti ratio in the deepest sediment layers was used as reference. the zr:ti ratio indicated the changes in grain size patterns and was used to understand the possible impacts of secular changes of hydrological dynamics the on the sediment quality (boyle, 2000). the ratio k:al was considered to identify mineral matter changes, since lower values generally reflect more-weathered mineral matter and viceversa (kauppila and salonen, 1997). all the stratigraphic plots were drawn with the software c2 version 1.7.2 (juggins, 2007). around 0.5 g of wet sediment of each subsample were extracted in 90% acetone overnight in the dark and under nitrogen atmosphere for photosynthetic pigments analysis, which were carried out at cnr-ise (verbania, italy). after centrifugation (3000 rpm, 10 min) the sediment was removed and the obtained extract was used to quantify total carotenoids (tcar) by double beam spectrophotometer (safas, uvmc2), and astaxanthin (asta) by reversed phase high-performance liquid chromatography using a thermo separation hplc (ultimate 3000). tcar was used to infer past total phosphorus concentration (car-tp) according to guilizzoni et al. (2011). subfossil cladocera cladocera remains were analyzed every fourth sample along the core following the methods described by szeroczyńska and sarmaja-korjonen (2007). about 2 cm3 of wet sediment were treated with koh (10%) and hcl (10%). asafranin-glicerol mixture was added to the cleaned subsamples of 0.1 ml in order to facilitate the identification of the remains under an optical microscope (leica dm2500) at 100-400x magnification. all cladocera remains (headshield, shell, postabdomen, postabdominal claws, mandible, caudal furca) were counted, and converted to number of individuals following frey (1986). taxonomical identification was based on flössner (2000), margaritora (1983) and szeroczyńska and sarmaja-korjonen (2007). three to six slides were counted for each sample in order to obtain a minimum of 100 cladocera individuals (kurek et al., 2010). this minimum was not achieved in a few samples with extremely scarce cladocera remains. in order to preliminary explore the influence of environmental variables on cladocera body size changes, ca. 30 carapace, mucro and antennules of different bosminidae species and 30 postabdominal claws of daphnidae were measured in each sample (korosi et al., 2008). environmental variables and data analysis homogenized monthly average air temperature data for the period 1870-2008 were obtained from the histalp webpage (2013) for the station torbole-riva del garda, which is located at 5 km away from lake ledro. in this work only annual and seasonal average air temperature were considered, as previous statistical analyses (milan, 2016) showed no relation between precipitation data and cladocera. information on lake nutrients and plankton were collected from the literature (casellato, 1990; boscaini et al., 2012) and from the environmental agency of the autonomous province of trento (unpublished data). past lake total phosphorus concentrations were inferred from both concentrations of total carotenoids (tcar, see above) and subfossil diatoms. diatom-inferred total phosphorus concentrations (ditp) were reconstructed basing on a weighted-average regressions (wa) with inverse deshrinking, calibrated against the nw-european datasets (nw-eu, bennion et al., 1996). further details on diatoms and pigment composition as well as tp reconstruction are available in milan (2016). the optimal partitioning method based on the sum of squares criterion and implemented in the zone software (lotter and juggins, 1991) was used to identify the homogenous cladocera zones along the master core, while the number of significant zones was established through comparison with the broken stick model (bennett, 1996). the binary logarithm-based shannon index (shannon and weaver, 1949) was applied to determine the diversity of fig. 1. study site including catchment area for lake ledro. dot indicates the coring point. no n c om me rci al us e o nly sediment cladocera and geochemistry of lake ledro 223 cladocera assemblages. the ratio of planktonic to littoral taxa was calculated for every analysed subsample, since planktonic species are expected to be dominant under warmer conditions and/or water level increase (sarmajakorjonen, 2001). bosmina longirostris (o.f. müller) was excluded from this classification due to its capacity to live in both the pelagic and littoral zones (szeroczyńska, 1998). specific ecological preferences were defined as in korhola (1990), frey (1986) and margaritora (1983). in order to identify patterns in the temporal evolution of cladocera assemblages, a non-metric multidimensional scaling (nmds, kruskal and wish, 1978) was applied to a bray & curtis dissimilarity matrix computed on the square root of cladocera abundances. the nmds analysis was performed with r 3.3.1 (r core team, 2016), vegan package version 2.4-1 (oksanen et al., 2016), and after 20 trials the nmds solution providing the lowest ‘stress’ (i.e. 0.16), which is measure of the configuration stability (legendre and legendre, 1998), was selected. a scree plot analysis was performed in order to determine the final number of nmds dimensions to be considered (legendre and legendre, loc. cit.). for the identification of the major environmental, climatic and geochemical drivers for the studied proxy, vector and surface fitting analyses were applied respectively to the sample scores of the nmds configurations. only variables significantly (p<0.05) correlated to sample score were considered in this work and plotted in the graph. the two analyses were computed using r 3.3.1 (r core team, 2016), vegan package version 2.4-1 (oksanen et al., 2016). a lowess interpolation of the non-contiguous radioisotopic ages was performed (r 3.3.1; r core team, loc. cit.) in order to attribute ages and sedimentation rates to each subsample within the 210pb dated core section. a kruskal-wallis test was applied as non-parametric test for non-normal distributed variables to sediments sections characterized by homogeneous clacodera body size, in order to verify the significance of different average sizes (kruskal and wallis, 1952). past software was used for the statistical analysis on the clacodera body measurements (hammer et al., 2010). results core chronology 210pb equilibrium was reached in the core of lake ledro at 28 cm depth, while 137cs showed a well resolved peak at 21.25 (coincident with a peak in 241am) and a minor one at 10.25 cm, corresponding to the fallout from atmospheric testing of nuclear weapons in 1963 and from the chernobyl accident in 1986, respectively. the chronologies and the sediment accumulation rates were calculated by the crs model using the 137cs peak at 21.25 cm as reference level for 1963. the sediment accumulation rates (fig. 2) showed a gradual increase in the last hundred years and a peak in the middle 1960s, which could be the result of either a sediment slumping or of the “century” flood event, which interested the north and central italy in november 1966 (malguzzi et al., 2006). the core showed major discontinuities at 40, 51, 59, 68 and 78 cm depth, which were interpreted as flood events by comparison with the structure of the core collected at 46 m depth in 2008 by simonneau et al. (2013b). other minor discontinuities detected in the core top 30 cm were related to minor floods (e.g., 1966) or to particularly rainy years (e.g., 1996, 2002, 2007). geochemistry and subfossil pigments wet density values oscillated in the core studied between 1.3 and 1.6 g cm-3. the highest values recorded between the core bottom and 28 cm (i.e., ad 1860±36) indicated high proportion of mineral matter during the little ice age (fig. 2). the upper 28 cm were characterized by gradually decreasing sediment densities with two peaks, one in the 1990s and one in the 2000s. water content oscillated around ~50% of fresh weight (fw) from the core bottom to 55 cm depth, and increased thereafter up to >80% in the upper 5 cm (fig. 2). organic content slightly decreased from core bottom to 28 cm depth (from 14% to 8% of dw), while it irregularly increased in the upper core section and reached maximum values in the early 1970s (at ~18 cm) and in the second half of the 2000s (at ~3 cm; fig. 2). the depth profile of astaxanthin (a marker for n-limited cyanobacteria and aquatic invertebrates, including zooplankton) oscillated between ~2 and 30 nmol g-1 loi throughout the core (fig. 2). total carotenoid concentration were low in the deeper core sections, while they increased up to 3.4 u g-1 loi during the 1980s and to ~4 u g-1 loi in the early 2000s (fig. 2). the past lake tp inferred on total carotenoid concentration (car-tp, fig. 2) oscillated between 5 and 8 µg l-1 from the core bottom up to ~20 cm depth (late 1960s), while the upper section showed two rapid increases up to mesotrophic level (>25 µg l-1), the first one in the 1970-1980s and the second one between the late 1990s to the beginning of the 2000s. these two maxima were separated by a sharp drop down to base line tp values in the 1990s (fig. 2). sediment surface car-tp concentrations (10-15 µg l-1) agreed with present mesotrophic status of lake ledro. the car-tp pattern at lake ledro was highly comparable to the tp profile inferred from subfossil diatoms (di-tp, fig. 2). however, di-tp concentrations reached in general higher values than car-tp and showed enhanced values also around 30 cm depth (i.e., during the middle 19th century). di-tp values oscillated around 10 no n c om me rci al us e o nly 224 m. milan et al. µg l-1from the bottom till the late 1960s (~20 cm depth), when they started to increase up to maximum values around 70 µg l-1 in the second half of the 1970s and the end of the 1980s, and to values of ~78 µg l-1 in the second half of the 2000s (fig. 2). subfossil cladocera the 33 cladocera taxa (6 planktonic and 27 littoral), which were identified in lake ledro belonged to the families leptodoridae, daphnidae, bosminidae, chydoridae, cercopagidae. three major zones were identified based on the cladocera assemblages and abundances (fig. 3). the first zone (lc1, 82.5-36.5 cm depth), was mainly characterized by the presence of b. longirostris, which was also responsible for the high values of total cladocera abundance (total) in the bottom core section. bosmina (e.) longispina leydig and a few individuals of chydoridae were also found in this section. the sediment layer around 38.5 cm depth revealed the lowest cladocera abundance value of the entire core, and in this core section only a few individuals of b. (e.) longispina, b. longirostris, acroperus harpae (baird) and alona affinis (leydig) were identified. in agreement with the species scarcity of this zone, the shannon index showed low values around 1. a rapid increase in littoral species, especially those connected with high water turbidity, such as alona rectangula sars and chydorus sphaericus (o.f. müller), or those associated with abundant detritus, like disparalona rostrata fig. 2. depth profiles of geochemical and biological proxies in the lake ledro. sed rate, sedimentation rate; wd, wet density (x-axis from 1 to 1.6); water, water content; om, organic content; asta, astaxanthin (loi, loss on ignition); tcar, total carotenoids concentrations; car-tp, total phosphorus concentration reconstructed from total carotenoids levels; di-tp, diatoms inferred total phosphorus concentrations. no n c om me rci al us e o nly sediment cladocera and geochemistry of lake ledro 225 (koch), characterized the zone lc2 (36.5-29.25 cm, fig. 3). high abundances of a. harpae and a. affinis were also found in this zone, while b. longirostris and b. (e.) longispina showed very low density in comparison to lc1. cladocera diversity reached the highest values of the entire in this section (shannon index = 3). the zone lc3 (29.25-0.0 cm) was characterized by an increase in density and diversity of planktonic species (fig. 3), which were used to define the two subzones lc3a and lc3b. lc3a (29.25-19.25 cm) showed a rapid decrease in total cladocera abundance, while only sporadic individuals of b. longirostris were identified. daphnia longispina o.f. müller increased gradually, and bosmina (e.) coregoni baird appeared for the first time in this section (i.e., after the 1930s). compared to the previous zone, the shannon index decreased in association to the abundance of planktonic species. the dominance of planktonic and the sporadic presence of littoral species marked zone lc3b (19.25-0.0 cm). planktonic species further increased since the early 1960s (19.25 cm) and culminated in the late 1980s, when they determined the peak in total cladocera abundance. a few individuals of b. longirostris were identified at the beginning of lc3b subzone, while it completely disappeared in the surface layers. the cladoceran predators bythotrephes longimanus leydig and leptodora kindtii (focke) appeared for the first time in this zone and reached their maximum values at 10.25 and 12.25 cm, respectively. cladocera diversity showed a further regular decrease since the beginning of the subzone lc3b (fig. 3). the highest cladocera abundance along the core largely corresponded to stages of higher astaxanthin concentrations (fig. 2). fig. 4 summarizes the results of the cladocera measurements. no trend was observed for d. longispina (not represented in fig. 4), while bosminidae exhibited a major shift during the 1980s. in particular, antennulae of b. (e.) coregoni clearly increased after the 1960s (~20 cm depth), and in particularly during the 1980s, while carapace and mucro sizes increased only since the 1980s (~10 cm depth) and following a decreasing stage from the 1960s to the 1980s (fig. 4). antennules and mucro of b. (e.) longispina showed a relatively comparable trend, while carapace size clearly increased only after the 1980s. despite the high abundance of b. longirostris especially fig. 3. depth profiles of: subfossil cladocera total abundance (total); ecological classification: planktonic (white), b. longirostris (grey), littoral (black); key species and shannon index. lc1-lc3b represented the homogeneous cladocera zones. species codes and full names of taxa are presented in supplementary tab. 2. no n c om me rci al us e o nly 226 m. milan et al. in the deeper core sections, the remains were often broken and the measurement of the body parts was hard. as major changes in bosmina body size occurred around ~10 cm depth, average sizes above and below this level were compared through the kruskal-wallys test (fig. 4, supplementary tab. 1). only the mucro of b. longispina were analyzed for three groups, i.e. between 0-10 cm, 10-26 cm and 26-82 cm. size differences between the two core sections were significant (p<0.05) for antennules and mucro of b.(e.) coregoni and for antennules and carapace of b.(e.) longispina. while differences between the three groups were significant (p<0.001) for mucro of b.(e.) longispina. environmental variables and data analyses the scree plot analysis indicated two dimensional nmds as sufficient to describe the diversity of subfossil cladocera data. the analysis revealed an evident grouping of the planktonic species in the nmds lower left quadrant, while littoral species were concentrated in the two upper quadrants. b. longirostris, resulted to be separated from all the other species in the lower right quadrant (fig. 5a), in agreement with its ecological characteristics (i.e., its capacity to live in both the pelagic and the littoral lake habitat). the separation of planktonic from littoral species along the first nmds dimension (dim1) corresponded to the species change which identified the separation of cladocera subzones lc3a and lc3b along the core (fig. 3). moreover, the sample score on the second nmds dimension (dim2, not shown) highlighted a drastic change at 38.5 cm depth, as already outlined by the cladocera assemblages (fig. 3). the vector fitting outlined strong positive relations of the cladocera assemblages with water content (water), tcar, car-tp and di-tp (0.30<r<0.69, p<0.001, fig. 5b), while presented a negative correlation with wd (r=0.62, p<0.001, fig. 5b). surface fitting confirmed the strong correlation between cladocera assemblages and di-tp, and the relation between planktonic species and high water tp levels (fig. 5b). the vector and the surface fitting analyses using temperature variables were limited only to the 210pb-dated core section. the vector fitting indicated only a weak correlation between cladocera composition and the histalp spring temperatures (tmm, r=0.45, p<0.05, fig. 5c) for the period between 1870 and 2008, while the surface fitting showed a correlation between mean annual air temperature and the recent samples (i.e., after the 1980). geochemical records (xrf data) the depth profiles of the majority of the analyzed elements and ratios showed continuous oscillations, with only some elements (such as mg and ti; fig. 6) showing a slight tendency to a reduction toward the sediment surface. mg and al oscillated around 2.5%, ca fluctuated around 21% and mn around 430 µg g-1. p and p:ti ratios exhibited stable values until ~25 and 20 cm depth, respectively (~1920s and 1950s), while both showed a slight increase and a peak at the beginning of the 2000s. in the same period br also recorded a peak. a generally declining trend was recorded for ti until the 1940s, when it decreased from values around 1400 µg g-1 to values around 1000 µg g-1. pb and pbef values increased during the first half of the 1900s until the peak in the 1970s. fluctuations were observed also in almost all the ratios (fig. 6). zr:ti showed an increase in the fluctuation frequency after the 1930-1940s. a marked increase in si:al values after the 1970s is superimposed to the long-term steady trend of this marker. in fig. 5d the vector fitting analysis showed a strong and positive relationship of the cladocera assemblages with different elements and ratios, i.e., p, cl, br, zr:ti, si:al, p:ti and pbef (0.25<r<0.3; p<0.05), while presented negative correlation with al, k, mn, ti and rb (0.24<r<0.37; p<0.05). surface fitting confirmed the strong correlation between cladocera assemblages and p:ti, and the relation between cladocera species and the changes in grain size (fig. 5d). discussion the paleolimnological study of lake ledro provided a retrospective analysis of the ecological and geochemical changes occurred in the last centuries in a small subalpine lake as a consequence of both anthropogenic and climate impacts. as highlighted by the 210pb-based chronology, the core layer at 27.75 cm depth was deposited 150±23 years ago. the core revealed several discontinuities in its structure, which were interpreted as flood records in relation to their aspect and enhanced wet density. the numerous flood markers in the deep section of the core investigated prevented the estimation of the core bottom age through the application of age-depths models. the arrangement of the flood records along the core and their higher frequency before the 1900s agreed with results of recent investigations focused on the hydrological history of lake ledro (simonneau et al., 2013a; vannière et al., 2013). no major discontinuities were found in the studied core after the 1900s. nevertheless, the enhanced sedimentation rates around 20 cm depth was interpreted as the century flood of 1966 (malguzzi et al., 2006), while the higher wet density and organic content registered in the surface layers could probably be identified as a result of the particularly abundant summer (as in 1996 and 2002) or winter (2004-2005, 2008-2009) precipitations (milan, 2016). the depth profile of astaxanthin showed high concentrations along the entire core and its peak values corresponded to maximum abundance of total cladocera no n c om me rci al us e o nly sediment cladocera and geochemistry of lake ledro 227 recorded in the late 1980s. on the other hand, tcar showed very low concentrations before the 1960s, suggesting low phytoplankton densities, possibly accompanied by poor preservation of the most labile carotenoids in the oldest core sections. the depth profile of the subfossil pigment-inferred lake tp concentrations (car-tp) reflected the tcar profile and its values were in general lower that tp concentrations inferred from sediment diatoms (di-tp), especially in the upper core section. both car-tp and di-tp showed stable oligotrophic conditions until the 1960s, and a successive increase up to maximum values during the 1980s and again at the beginning of the 2000s. however, car-tp concentration never exceeded mesotrophic values (~26 mg l-1), while di-tp reached clearly eutrophic levels (~78 mgl-1) in the same periods. maximum di-tp levels agree with the lake eutrophic condition reported in the literature (casellato, 1990; boscaini et al., 2012). the discrepancy between tp reconstruction based on subfossil pigments and diatoms, respectively, especially for periods of higher lake productivity, has been often observed and may be related to the higher degradation rates of pigments in the sediment respect to diatoms (guilizzoni et al., 2011). the peak di-tp values before 1860s were not matched by correspondent maximum in fig. 4. b.(e.) coregoni (a-c) and b.(e.) longispina (d-f) body size measurements: antennules (a,d), carapace (b,e), mucro (c,f). the values indicate the average deviations of each sample from the general average size calculated for all the individuals measured (bold horizontal line). x-axis refers to core depths, from core surface (0 cm) down to core bottom (82.5 cm). no n c om me rci al us e o nly 228 m. milan et al. fig. 5. cladocera non-metric multidimensional scaling (nmds) ordination. a) cladocera species distributions on the samples in the ordination space; numbers in the plots refer to core samples of lake ledro; species codes and full names of taxa are presented in supplementary tab. 2. ellipses show the major group of zooplankton species. b) vector and surface fitting for environmental variables; the orientations of the vectors show the change direction for each environmental variable and the length indicate the correlation between the environmental variable and the sample ordination; in the plot are shown only the results that exhibited significant (p<0.05) correlations; wd, wet density; water:, water content; om, organic content; asta, astaxanthin; tcar, total carotenoids concentrations; cartp, total phosphorus concentration reconstructed from total carotenoids levels; di-tp, diatoms inferred total phosphorus concentrations; the surface fitting is based on correlations with di-tp, and the numbers on the surface lines represent the di-tp values. c) vector and surface fitting for climate variables. only tmm (spring air temperature from march to may) exhibited a significant correlation with p<0.05; the surface fitting is based on correlations with the mean annual air temperature and the numbers on the surface lines are measured mean annual air temperatures. d) vector and surface fitting for geochemical variables based on xrf analysis. in the plot are shown only the results that exhibited significant (p<0.05) correlations; pdef, pb enrichment factors; the surface fitting is based on correlations with the p:ti and the numbers on the surface lines are measured p:ti ratio. no n c om me rci al us e o nly sediment cladocera and geochemistry of lake ledro 229 car-tp. this discrepancy, which was due to a stage of very low abundances of planktonic diatoms and a high share of meso-to eutraphentic benthic diatoms, was interpreted as an effect of inorganic and organic materials dragged from the catchment into the lake by a major flood event occurred in the first half of 1800 (milan, 2016). this may have induced a persistent increase in water turbidity and the consequent depression of phytoplankton productivity. the presence of cladocera species preferring turbid waters, such as a. rectangula and c. spahericus supported the hypothesis of low water transparency (margaritora, 1983), while the presence of b. longirostris during and after this flood confirmed the increased in nutrient concentrations (boucherle and züllig, 1983). the car-tp and di-tp increase after the 1960 and their maximum values during the 1980s corresponded to the general trophic evolution observed in other alpine lakes in relation to the post-war economic development (berthon et al., 2013; milan et al., 2015). the tp pulse during the 1980s was confirmed by the limnological data collected from the autonomus province of trento (unpublished data). the peak at the beginning of the 2000s was also recorded in other alpine lakes (milan et al., 2015; berthon et al., 2013; tolotti et al., 2012) and could be related to the particular weather conditions, especially to the high summer temperatures and the high snow precipitations, registered during the first half of the 2000s. in fact, simonneau et al. (2013b) demonstrated that snow thaw is particularly efficient in dragging the nutrients stored in the lake catchment down to the lake ecosystems. although cladocera communities were not always well preserved in the sediments of lake ledro, especially during the flood events, both littoral and planktonic taxa were well represented in the core studied. the deeper section, which was supposed to cover the little ice age (lia), was characterized by an unexpected high abundance of total cladocera as a consequence of high densities of b. longirostris and b.(e.) longispina. usually, the presence of bosminidae, and in particular of b. longirostris, indicate high levels of lake nutrient concentrations (hofmann, 1998; korosi et al., 2013), which contrasts with the information provided by the tp reconstruction based on subfossil pigments and diatoms. however, the higher density of b. longirostris is not in contrast with the higher hydrofig. 6. selected geochemical depth profiles. geochemical elements are divided in major groups: major, redox, lithogenic and trace. pbef, pb enrichment factors. no n c om me rci al us e o nly 230 m. milan et al. logical variability of the pre-1860 period and could be related to higher amount of particulate material and nutrients entering into the lake from the catchment area during flood events (manca et al., 2007). on the other hand, the presence of a. harpae and a. affinis in the core bottom confirmed the deposition of these sediment layers during the lia, as these two species were considered as early immigrants after ice retreat (kamenik et al. 2007) and are classified as “arctic” and “sub-arctic” species, respectively (harmsworth, 1968). in addition, a. harpae and a. affinis were the sole species identified at 38.5 cm depth (ca. in the early 1800s), which reinforces the hypothesis that this sediment section was deposited during the coldest period recorded in the alpine region (de jong et al. 2013). the lowest total cladocera abundance and the particular species composition recorded in this layer agreed also with what observed in lake garda at the beginning of the 1800s. in particular, the core collected from the deepest lake basin showed the lowest cladocera density and exclusive the presence of the a. harpae during this particularly cold period (milan, 2016). cladocera community changed completely its species composition and reached its highest richness during the first half of the 1800 (zone lc2), which was interested by one of the biggest flood event in the whole record. bosminidae were substituted by species preferring turbid waters, such as a. rectangula and c. sphaericus, and/or associate with detritus or sand, such as d. rostrata and monospilus dispar (sars). the presence of campotercus rectirostris schödler and pleuroxus spp. (baird), not showed in fig. 3, suggested also an increase in mean water temperatures. the decrease in littoral species and the disappearance of b. longirostris in zone lc3a suggested a lake recovery to pre-flood event status and low tp concentrations. the decrease in nutrient was confirmed by the di-tp, which showed a return down to oligotrophic levels after the flood. the dominance of b.(e.) longispina, d. longispina and b. longirostris since the 1960s indicated a nutrient enrichment stage (boucherle and züllig, 1983), in agreement with the independent pigmentand diatom-inferred tp. the planktonic species reached their highest density during the 1980s, i.e. in concomitance with the maximum car-tp and di-tp values. the tp peak of 2005 and the following decrease in nutrients, did not correspond to evident changes in cladocera assemblages, probably in relation to resuspension of cladocera remains in uppermost core layers (szeroczyńska, personal communication). the multivariate statistical analyses outlined that secular changes in cladocera communities of lake ledro responded weakly to mean air temperature changes, while the response to hydrological events, especially to the major flood occurred in the mid-1800s, appeared more pronounced. on the contrary, lake nutrient levels emerged as the principal cladocera driver after the 1960s. in fact significant correlations were found between cladocera species (in particular those identified in the upper layers) and lake nutrient level, as well as geochemical elements and ratios, while correlations with temperature variables were scarce. the highly comparable gradients of environmental, climate and geochemical variables in the vector and surface fitting analyses suggested a simultaneous and overlapped role of all the three variable groups considered in the present study in driving changes in the cladocera assemblage. this strongly hinders the discrimination and isolation of effects by single variables. on the other hand, this study confirmed the response of cladocera assemblages to extreme events, which were able to induce major habitat changes, as similarly emerged from the paleolimnological study of lake garda (milan, 2016). since cladocera occupy an intermediate level between top-down regulators (fish and invertebrate predators) and bottom-up factors (nutrient and phytoplankton, jeppesen et al., 2001; szeroczyńska, 2006), the role of predation pressure should also be considered when evaluating long term changes in the cladocera composition. however, as only sparse fish information are available for lake ledro, the analysis of size structure of daphnia spp. and bosmina spp. was used as an indirect proxy for the reconstruction of food web changes. in fact, previous studies could related the decrease in relative abundance and in body size of daphnia spp. and bosmina spp. to increases in fish predation (korosi and smol, 2012). on the other hand, bosmina spp. specimens with long antennules and mucro are usually recorded in conditions of high predation by invertebrates and low fish predation (korosi and smol, loc. cit.). although in lake ledro daphnia spp. did not show changes in the body size structure, bosmina spp. showed a size increase of the different measured body parts since the second half of the 1980s. this alteration was linked to the presence of bythotrephes longimanus leydig and leptodora kindtii (focke), which appeared in lake ledro after the 1960s and reached their maximum densities in the second half of the 1980s. changes in cladocera body size in relation to enhanced invertebrate predation were observed also in lake maggiore since the 1980s (manca et al., 2007). the secondary size shift observed during the 1930s could not be related to the invertebrate pressure since b. longimanus and l. kindtii were not recorded in the sediment layer deposited in that period, in agreement with the first record of both species in the lake waters only after the 1950s (barbato, 1977), while the correlation with other invertebrates predators could not be proved, due to the lack of such information. nevertheless, changes in cladocera body size are known to depend from several limnological variables, such as the tp concentrations. in fact, larger cladocera prevail in lakes with low tp values, as they are characterized by no n c om me rci al us e o nly sediment cladocera and geochemistry of lake ledro 231 lower limiting thresholds for nutrients in comparison to smaller cladocera (korosi et al., 2008). in lake ledro the secondary increase in cladocera body size might be related to lower nutrient concentrations during the 1930s, as outlined by pigmentand diatom-inferred tp reconstructions. although cladocera of lake ledro revealed a strong response to the major flood event in the early 1800, no clear responses were recorded to minor hydrological perturbations as well as to lake-level regulations after 1929. the geochemical analysis showed continuous fluctuations of elements and ratios throughout the entire core in response to the constant influence of the large catchment area on the lake system, but they did not showed major changes after the lake regulation. the depth profile of pb concentrations showed the typical trend recorded in the alpine region and confirmed the larger scale atmospheric deposition pattern on the southern alps (rogora et al., 2006; thevenon et al., 2011). the pb pattern of lake ledro was highly comparable with that observed in lake garda (milan, 2016), likely in relation to the vicinity of the two lakes. the mn:fe ratios were considered in order to assess the historical redox conditions, as anoxic conditions of the hypolimnion of lake ledro were recorded during the 1970s-1980s (casellato, 1990). however, no major long term changes in redox conditions were observed, possibly as seasonal oxygen depletion at sediment-water interface represent a constant feature of this lake (boscaini et al., 2012). si:al ratios were calculated in order to infer biogenic silica, which in turn are considered as a proxy of totaldiatom production (peinerud et al., 2001). the si:al profile is highly comparable with the temporal variability of large colonial pennate diatoms along the core (r=0.75, p<0.001; milan, 2016). in contrast, si:al ratios and the total diatom concentrations presented comparable trends only after the 1960s (r=0.62, p<0.001; milan, 2016), which might be explained by the different silicization degree of different diatom taxa (peinerud et al., 2001). the p:ti ratios agreed with the nutrients increase after the 1960s suggested by the tp reconstruction, while the peak in p:ti ratio at the beginning of the 2000s agree with the increase in lake productivity indicated by the high diatom concentrations and subfossil pigments concentrations (milan, 2016). the proxy for changes in sediment grain size, i.e. zr:ti ratios, recorded continuous fluctuations since the 1940s, possibly in relation to the lake-level regulation started in 1929. the rb depth profile showed a comparable trend respect to zr:ti ratios, and changes in rb content could be related to the lake-level regulation during the hydroelectric activity as observed in the study on lake majeur (simonneau et al., 2013a). on the other hand the k:al ratios, which is a proxy of the mineral matter quality, registered only some changes during the flood events, while no significant changes were recorded after the 1940s. the xrf analyses on the sediment core of lake ledro were able to track secular changes in geochemical elements and outlined also a certain influence of lakelevel regulations on the sediment properties. conclusions studies of sediment cladocera remains are still scattered for subalpine lakes. sediments of lake ledro were more intensively studied for hydrological purposes than for ecological reconstructions, in relation to its peculiar morphological and hydrological features. the present paleolimnological study aimed at providing a general overview of long term effects of climatic, anthropogenic and hydrological impacts on environment and ecology of lake ledro, a small subalpine lake characterized by a large catchment area. subfossil cladocera of lake ledro confirmed the information on ecological preferences of several species, which were previously identified in lakes located in different european regions and the large and deep lake garda. cladocera of lake ledro revealed a good capability in tracking major climate related changes, such as major increases or decrease in water temperature, as during the little ice age, and major events of hydrological variability (i.e. floods). the development of new habitats during and after major floods in relation to higher inputs of material from the lake catchment resulted in clear changes in cladocera species assemblages and diversity. in addition, the drastic changes in species composition, together with the increase in dominance of planktonic taxa after the 1960s, confirmed the effects of nutrient enrichment on cladoceran communities observed in the majority of the alpine lakes. similarly to what observed in lake garda, the response of subfossil cladocera of lake ledro to water temperature was stronger during periods of low lake nutrient levels, while response to nutrients prevailed in nutrient enriched stages. the preliminary study of changes in cladocera body size suggested the key role of invertebrate predators in affecting bosminidae. this corroborated the usefulness of such analysis in tracking food web changes when information on fish population is not available. the analyses of subfossil cladocera and geochemical proxies allowed to outline a weak role of lake level regulation in affecting lake habitat and ecology. nevertheless, this work outlined thatthe combined analysis of biological and geochemical sediment proxies is able to track the role hydrological variability in affecting environmental and ecological lake conditions. no n c om me rci al us e o nly 232 m. milan et al. acknowledgments the authors thank adriano boscaini, flavia brescancin (e. mach foundation istituto agrario di s. michele all’adige) and johan rydberg (umeå university) for support in the field and in the laboratory. krystyna szeroczyńska, joanna stańczak and elżbieta kowalczyk (institute of geological sciences, polish academy of sciences in warsaw) for their support in laboratory and with cladocera identification. the authors are grateful to andrea lami (ise national research council) and to handong yang (ensis ltd. university college of london-uk) for pigment and radiometric analyses, respectively. authors thanks rosalie bruel and a second anonymous reviewer for their constructive comments. the first author thanks the international research school in applied ecology (irsae) for the cofounding. references alefs j, müller j, 1999. differences in the eutrophication dynamics of ammersee and starnberger see (southern germany), reflected by the diatom succession in varve-dated sediments. j. paleolimnol. 21:395-407. appleby pg, 2001. basin analysis, coring, and chronological techniques, pp. 171-203. in: w.m. last and j.p. smol (eds.), tracking environmental change using lake sediments. kluwer, dordrecht. barbato g, 1977. note idrobiologiche sul lago di ledro. ann. mus. civ. st. nat. 14:92-109. battarbee rw, 2000. palaeolimnological approaches to climate change, with special regard to the biological record. quaternary sci. rev. 19:107-124. bennett kd, 1996. determination of the number of zones in a biostratigraphical sequence. new phytol. 132:155-170. bennion h, davidson ta, sayer cd, simpson g, rose nl, sadler jo, 2015. harnessing the potential of the multi-indicator palaeolimnological approach: an assessment of the nature and causes of ecological changes in a eutrophic shallow lake. freshwater biol. 60:423-1442. bennion h, juggins s, anderson nj, 1996. predicting epilimnetic phosphorus concentrations using an improved diatombased transfer function and its application to lake eutrophication management. environ. sci. technol. 30:2004-2007. berthon v, marchetto a, rimet f, dormia e, jenny jp, pignol c, perga me, 2013. trophic history of french sub-alpine lakes over the last ~150 years: phosphorus reconstruction and assessment of taphonomic biases. j. limnol. 72:417-429. boscaini a, brescancin f, salmaso n, 2012. [progetto di ricerca per lo studio dei fattori chimico-fisici che regolano lo sviluppo del ciano batterio planktothrix rubescens nel lago di ledro].[report in italian]. e. mach foundation, s. michele all’adige: 67 pp. boyle jf, 2000. rapid elemental analysis of sediment samples by isotope source xrf. j. paleolimnol. 23:213-221. boucherle mm, züllig h, 1983. cladoceran remains as evidence of change in trophic state in three swiss lakes. hydrobiologia 103:141-146. casellato s, 1990. [il lago di ledro, valutazione del suo stato trofico].[report in italian]. provincia autonoma di trento: 111 pp. davidson ta, sayer c, perrow m, bramm m, jeppesen e, 2007. are the controls of species composition similar for contemporary and sub-fossil cladoceran assemblages? a study of 39 shallow lakes of contrasting trophic status. j. paleolimnol. 38:117-134. de jong r, kamenik c, grosjean m, 2013. cold-season temperatures in the european alps during the past millennium: variability, seasonality and recent trends. quaternary sci. rev. 82:1-12. eggermont, h, martens k, 2011. preface: cladocera crustaceans: sentinels of environmental change. hydrobiologia 676:1-7. finney bp, gregory-eaves i, sweetman j, douglas ms, smol jp, 2000. impacts of climatic change and fishing on pacific salmon abundance over the past 300 years. science 290:795799. flössner d, 2000. [die haplopoda und cladocera (ohne bosminidae) mitteleuropas].[book in german]. backhuys: 428 pp. frey d, 1986. cladocera analysis, p. 677-692. in: b.e. berglund (ed.), handbook of holocene palaeoecology and plaeohydrology. citeseer garcia-ruiz jm, 2010. the effects of land uses on soil erosion in spain: a review. catena 81:1-11. guilizzoni p, marchetto a, lami a, gerli s, musazzi s, 2011. use of sedimentary pigments to infer past phosphorus concentration in lakes. j. paleolimnol. 45: 433-445. guilizzoni, p, lami a, manca m, musazzi s, marchetto a, 2006. palaeoenvironmental changes inferred from biological remains in short lake sediment cores from the central alps and dolomites. hydrobiologia 562:167-191. hammer ø, harper dat, ryan pd, 2001. past-paleontological statistics software package for education and data analysis. palaeontol. electronica 4:9. harmsworth rv, 1968. the developmental history of blelham tarn (england) as shown by animal microfossils, with special reference to the cladocera. ecol. monogr. 38:223-241. histalp, 2013. long series subset. available from: http://www.zamg.ac.at/histalp/content/view/35/1/index.html hofmann w, 1998. the response of bosmina (eubosmina) to eutrophication of upper lake constance: the subfossil record, p. 275-283. in: e. bäuerle and u. gaedke (eds.), lake constance: characterization of an ecosystem in transition. e. schweizerbart’sche verlagsbuchhandlung. jeppesen e, leavitt p, de meester l, jensen jp, 2001. functional ecology and palaeolimnology: using cladoceran remains to reconstruct anthropogenic impact. trends ecol. evol. 16:191-198. jeziorski a, yan nd, paterson am, desellas am, turner ma, jeffries ds, keller b, weeber rc, mcnicol dk, palmer me, 2008. the widespread threat of calcium decline in fresh waters. science 322:1374-1377. no n c om me rci al us e o nly sediment cladocera and geochemistry of lake ledro 233 juggins s, 2007. c2 version 1.5 user guide. software for ecological and palaeoecological data analysis and visualisation. newcastle university, newcastle upon tyne, uk 73. kamenik c, szeroczyńska k, schmidt r, 2007. relationships among recent alpine cladocera remains and their environment: implications for climate-change studies. hydrobiologia 594:33-46. kauppila t, salonen vp, 1997. the effect of holocene treeline fluctuations on the sediment chemistry of lake kilpisjarvi, finland. j. paleolimnol. 18:145-163. korhola a, 1990. paleolimnology and hydroseral development of the kotasuo bog, southern finland, with special reference to the cladocera. suomalainentiedeakatemia. korhola a, tikkanen m, weckström j, 2005. quantification of holocene lake-level changes in finnish lapland using a cladocera-lake depth transfer model. j. paleolimnol. 34:175190. korosi jb, kurek j, smol jp, 2013. a review on utilizing bosmina size structure archived in lake sediments to infer historic shifts in predation regimes. j. pl. res. 35:444-460. korosi, jb, paterson am, desellas am, 2008. linking mean body size of pelagic cladocera to environmental variables in precambrian shield lakes: a paleolimnological approach. j. limnol. 67:22-34. korosi, jb, paterson am, desellas am, smol jp, 2010. a comparison of pre-industrial and present-day changes in bosmina and daphnia size structure from soft-water ontario lakes. can. j. fish. aquat. sci. 67:754-762. korosi, jb, smol jp, 2012. examining the effects of climate change, acidic deposition, and copper sulphate poisoning on long-term changes in cladoceran assemblages. aquat. sci. 74:781-792. kruskal jb, wish m, 1978. multidimensional scaling. sage publ., beverly hills: 93 pp. kruskal wh, wallis wa, 1952. use of ranks in one-criterion variance analysis. j. am. stat. assoc. 47: 583-621. kurek j, korosi jb, jeziorski a, smol jp, 2010. establishing reliable minimum count sizes for cladoceran subfossils sampled from lake sediments. j. paleolimnol. 44:603-612. legendre pl, legendre l, 1998. numerical ecology. 2nd ed. elsevier: 300 pp. lotter a, juggins s, 1991. polprof, tran and zone: programs for plotting, editing and zoning pollen and diatom data. inqua subcommission for the study of the holocene working group on data-handling methods. newsletter 6:4-6. magny m, joannin s, galop d, vannière b, haas jn, bassetti m, bellintani p, scandolari r, desmet m, 201. holocene palaeohydrological changes in the northern mediterranean borderlands as reflected by the lake-level record of lake ledro, northeastern italy. quaternary res. 77:382-396. malguzzi p, grossi g, buzzi a, ranzi r, buizza r, 2006. the 1966 “century” flood in italy: a meteorological and hydrological revisitation. j. geoph. res. 111:d24. manca m, torretta b, comoli p, amsinck sl, jeppesen e, 2007. major changes in trophic dynamics in large, deep sub-alpine lake maggiore from 1940s to 2002: a high resolution comparative palaeo-neolimnological study. freshwater biol. 52:2256-2269. marchetto a, lami a, musazzi s, massaferro j, langone l, guilizzoni p, 2004. lake maggiore (n. italy) trophic history: fossil diatom, plant pigments, and chironomids, and comparison with long-term limnological data. quatern. int. 113:97-110. margaritora f, 1983. [cladoceri (crustacea: cladocera) guide per il riconoscimento delle specie animali delle acque interne italiane].[report in italian]. consiglio nazionale delle ricerche, rome. martin-puertas c, jimenez-espejo f, martinez-ruiz f, nietomoreno v, rodrigo m, mata mp, valero-garces bl, 2010. late holocene climate variability in the southwestern mediterranean region: an integrated marine and terrestrial geochemical approach. clim. past 6:807-816. milan m, 2016. long-term development of subalpine lakes: effects of nutrients, climate and hydrological variability as assessed by biological and geochemical sediment proxies. ph.d. thesis, umeå university. milan m, bigler c, salmaso n, guella g, tolotti, 2015. multiproxy reconstruction of a large and deep subalpine lake’s ecological history since the middle ages. j. great lakes res. 41:982-994. nevalainen l, 2012. distribution of benthic microcrustaceans along a water depth gradient in an austrian alpine lake-sedimentary evidence for niche separation. limnologica 42:65-71. oksanen j, blanchet fg, friendly m, kindt r, legendre p, mcglinn d, minchin pr, o’hara rb, simpson gl, solymos p, stevens mhh, szoecs e, wagner h. 2016. package ‘vegan’. community ecology package, version: 2.4.1. peinerud ek, ingri j, ponter c, 2001. non-detrital si concentrations as an estimate of diatom concentrations in lake sediments and suspended material. chem. geol. 177:229-239. perga me, frossard v, jenny jp, alric b, arnaud f, berthon v, black jl, domaizon i, giguet-covex c, kirkham a, magny m, manca m, marchetto a, millet l, paillès c, pignol c, poulenard j, reyss jl, rimet f, sabatim p, 2015. high-resolution paleolimnology opens new management perspectives for lakes adaptation to climate warming. front. ecol. evol. 3:1-17. r core team, 2016. r: a language and environment for statistical computing. r foundation for statistical computing, vienna, austria. url https://www.r-project.org/ rogora m, mosello r, arisci s, brizzio m, barbieri a, balestrini r, waldner p, schmitt m, stahli m, thimonier a, kalina m, puxbaum h, nickus u, ulrich e, probst a, 2006. an overview of atmospheric deposition chemistry over the alps: present status and long-term trends. hydrobiologia 562:17-40. rosén p, vogel h, cunningham l, reuss n, conley dj, persson p, 2010. fourier transform infrared spectroscopy, a new method for rapid determination of total organic and inorganic carbon and biogenic silica concentration in lake sediments. j. paleolimnol. 43:247-259. rydberg j, 2014. wavelength dispersive x-ray fluorescence spectroscopy as a fast, non-destructive and cost-effective analytical method for determining the geochemical composition of small loose-powder sediment samples. j. paleolimnol. 52:265-276. sarmaja-korjonen k, 2001. correlation of fluctuations in cladoceran planktonic: littoral ratio between three cores from a small lake in southern finland: holocene water-level changes. holocene 11:53-63. no n c om me rci al us e o nly 234 m. milan et al. shannon ce, weaver w, 1949. the mathematical theory of communication. university of illinois press: 144 pp. simonneau a, chapron e, courp t, tachikawa k, le roux g, baron s, galop d, garcia m, di giovanni c, motellicaheino m, mazier f, foucher a, houet t, desmet m, bard e, 2013a. recent climatic and anthropogenic imprints on lacustrine systems in the pyrenean mountains inferred from minerogenic and organic clastic supply (vicdessos valley, pyrenees, france). holocene 23:1764-1777. simonneau a, chapron e, vannière b, wirth sb, gilli a, di giovanni c, anselmetti fs, desmet m, magny m, 2013b. mass-movement and flood-induced deposits in lake ledro, southern alps, italy: implications for holocene palaeohydrology and natural hazards. clim. past. 9:825-840. szeroczyńska k, 1998. palaeolimnological investigations in poland based on cladocera (crustacea). palaeogeogr. palaeoclimatol. palaeoecol. 140:335-345. szeroczyńska k, sarmaja-korjonen k, 2007. atlas of subfossil cladocera from central and northern europe. friends of the lower vistula society: 84 pp. szeroczyńska k, 2006. the significance of subfossil cladocera in stratigraphy of late glacial and holocene. studia quaternaria 23:37-45. thevenon f, guedron s, chiaradia m, loizeau jl, pote j, 2011. (pre-) historic changes in natural and anthropogenic heavy metals deposition inferred from two contrasting swiss alpine lakes. quat. sci. rev. 30:224-233. tolotti m, thies h, nickus u, psenner r, 2012. temperature modulated effects of nutrients on phytoplankton changes in a mountain lake. hydrobiologia 698:61-75. vannière b, magny m, joannin s, simonneau a, wirth sb, hamann y, chapron e, gilli a, desmet m, anselmetti fs, 2013. orbital changes, variation in solar activity and increate anthropogenic activities: controls on the holocene flood frequency in the lake ledro area, northern italy. clim. past. 9: 1193-1209. wessels m, mohaupt k, kümmerlin r, lenhard a, 1999. reconstructing past eutrophication trends from diatoms and biogenic silica in the sediment and the pelagic zone of lake constance, germany. j. paleolimnol. 21:171-192. wunsam s, schmidt r, klee r, 1995. cyclotella-taxa (bacillariophyceae) in lakes of the alpine region and their relationship to environmental variables. aquat. sci. 57:1015-1621. no n c om me rci al us e o nly layout 1 introduction saxitoxins or paralytic shellfish toxins (psts) are a group of potent alkaloid neurotoxins produced by a number of freshwater cyanobacterial species, as well as by certain marine dinoflagellates. paralytic shellfish toxins accumulate in the food chain, and consumption of toxincontaminated seafood can lead to potentially severe poisoning. health risks are also associated with the use of water sources harbouring pst-producing cyanobacteria for e.g. drinking water production or recreation. in europe, psts have been detected widely in freshwater lakes, rivers and reservoirs; in czech republic (jančula et al., 2014) and spain pst production has been associated with blooms of aphanizomenon spp. (wörmer et al., 2011). in finland (rapala et al., 2005; savela et al., 2015) and denmark (kaas and henriksen, 2000) dolichospermum lemmermannii (richter) p. wacklin, l. hoffmann & j. komárek (formerly anabaena lemmermannii, wacklin et al., 2009), a species closely related to aphanizomenon gracile lemmermann (cirés and ballot, 2016), has been implicated as the most likely pst-producing species. moreover, biosynthesis of psts has been confirmed in numerous strains of a. gracile from germany, portugal and france (pereira et al., 2004; ballot et al., 2010; ledreux et al., 2010; casero et al., 2014; cires et al., 2014) and cuspidothrix issatschenkoi (usachev) p. rajaniemi, komárek, r. willame, p. hrouzek, k. kastovská, l. hoffmann & k. sivonen (formerly aphanizomenon issatschenkoi, rajaniemi et al., 2005) from portugal advances in oceanography and limnology, 2017; 8(1): 61-70 article doi: 10.4081/aiol.2017.6319 this work is licensed under a creative commons attribution-noncommercial 4.0 international license (cc by-nc 4.0). first report of cyanobacterial paralytic shellfish toxin biosynthesis genes and paralytic shellfish toxin production in polish freshwater lakes henna savela,1* lisa spoof,2 niina höysniemi,1 markus vehniäinen,1 joanna mankiewicz-boczek,3 tomasz jurczak,4 mikolaj kokociński,5 jussi meriluoto2 1biotechnology, department of biochemistry, university of turku, tykistökatu 6a, 20520 turku, finland; 2department of biochemistry and pharmacy, faculty of science and engineering, åbo akademi university, tykistökatu 6a, 20520 turku, finland; 3european regional centre of ecohydrology of the polish academy of sciences, tylna 3, 90-364 łódź, poland; 4department of applied ecology, faculty of biology and environmental protection, university of łódź, banacha 12/16, 90-237 łódź, poland; 5department of hydrobiology, faculty of biology, adam mickiewicz university, umultowska 89, 61-614 poznań, poland *corresponding author. e-mail: henna.savela@utu.fi abstract in central and southern europe, aphanizomenon spp., a. gracile lemmermann in particular, have been associated with paralytic shellfish toxin (pst) production. in western poland, a. gracile is very common, and cylindrospermopsis raciborskii (woloszyńska) seenayya & subba raju, another potentially pst-producing species, is often found as well. to date it is, however, unknown if the cyanobacterial populations in this area harbour the genetic capability to produce psts, and to what extent toxin biosynthesis occurs. the objective of this study was to survey the prevalence of potentially pst-producing cyanobacteria by measuring paralytic shellfish toxin biosynthesis gene sxtb copy numbers, sxta, sxtg and sxts gene presence, and pst concentrations in polish freshwater lakes. in total, 34 lakes in western poland were sampled twice during summer 2010. the presence of pst biosynthesis genes sxta, sxtg and sxts was determined using conventional qualitative pcr. quantitative pcr (qpcr) was used to measure sxtb copy numbers, and the samples were analysed for psts using ion-pair high performance liquid chromatography with post-column oxidation and fluorescence detection (hplc-fld). cyanobacteria carrying the sxtb gene were present in 23.5% of all samples (n=16) and in 14 lakes of the studied 34. gene copy numbers ranged from 8.2×104 to 5.1×107 sxtb copies l–1 (mean 3.8×106). the median was 4.5×105 sxtb gene copies l–1 and the majority of results clustered at the lower end of the sxtb qpcr linear range. in 12 out of the 16 samples positive for sxtb the gene co-occurred with the other three targeted pst biosynthesis genes sxta, sxtg and sxts. however, five additional samples lacked one or two of the targeted four genes. thirteen samples contained psts, of which 12 samples at levels <0.072 µg l–1, i.e., close to or below the quantitative detection limit of the hplc-fld method (0.01 µg l–1). one sample contained 0.57 µg l–1 saxitoxin, co-occurring with all four sxt genes studied. no correlation between pst and sxt gene occurrence or copy numbers was observed. a. gracile and c. raciborskii occurred in 92% and 50% of samples, respectively, containing psts, sxt genes or both. in conclusion, the results confirm that potential pst producers constitute an established subpopulation of cyanobacteria in polish freshwater lakes. however, none of the sxt genes targeted in this study could serve as a reliable marker for active pst biosynthesis. key words: saxitoxin; paralytic shellfish toxins; cyanobacteria; lakes; quantitative pcr; pcr. received: 30 september 2016. accepted: 30 january 2017.no nco mm er cia l u se on ly h. savela et al.62 (pereira et al., 2000; li et al., 2003). other pst-producing cyanobacteria encountered elsewhere in the world include cylindrospermopsis raciborskii (woloszyńska) seenayya & subba raju (lagos et al., 1999), dolichospermum circinale (rabenhorst ex bornet & flahault) p. wacklin, l. hoffmann & j. komárek (formerly anabaena circinalis, wacklin et al., 2009) and plectonema wollei farlow ex gomont (carmichael et al., 1997), raphidiopsis brookii hill (yunes et al., 2009), scytonema sp. (smith et al., 2011), geitlerinema spp., phormidium uncinatum gomont ex gomont, and cylindrospermum stagnale bornet & flahault (borges et al., 2015). of the above mentioned species, c. raciborskii is commonly found also in european lakes, including poland (kokociński and soininen, 2012), but has not been associated with pst production in europe to date. psts affect vertebrate nerve function by reversibly blocking voltage-gated ion channels, primarily na+ (llewellyn, 2006), but also k+ and ca2+ channels are affected (kao, 1986, wang et al., 2003, su et al., 2004). ion channel blockage leads to inhibition of neuronal signal transduction, the extent of which is directly dependent on pst concentration. disruption of neuronal signals leads to symptoms ranging from slight numbness and tingling of extremities to full respiratory arrest and death in extreme cases (llewellyn, 2006). the effectivity of ion channel blockage depends on the structural characteristics of each given pst variant, 57 of which have been identified (wiese et al., 2010). the biosynthesis of all psts is directed by a suite of biosynthetic genes (sxt), first described in c. raciborskii strain t3 (kellmann et al., 2008). since then, toxinspecific molecular genetic analysis has been employed in the study of pst-producing cyanobacteria in culture and in the field. several studies on european cyanobacteria have been carried out on strains isolated from freshwater lakes and reservoirs, and many have targeted the pst biosynthesis gene sxta showing that while all pst-producing strains do carry the gene, it can also be found in anabaena, anabaenopsis and aphanizomenon strains which appear to lack the ability to produce the toxins (ballot et al., 2010; ledreux et al., 2010; cires et al., 2014). other sxt genes, including sxtg, sxtb, sxti and sxtx, have been studied to a lesser extent (casero et al., 2014, gkelis and zaoutsos, 2014, savela et al., 2015). only a few studies have investigated the co-occurrence of sxt genes and psts in environmental samples in europe (gkelis and zaoutsos, 2014; savela et al., 2015) or elsewhere in the world (al-tebrineh et al., 2012; bowling et al., 2013). cyanobacterial toxins are common in western polish lakes; cylindrospermopsin, produced primarily by a. gracile, is found at concentrations exceeding who recommendations (kokociński et al., 2013). hepatotoxic microcystins are another common cyanotoxin group in polish lakes (mankiewicz-boczek et al., 2006). however, neither the prevalence of potentially pst-producing cyanobacteria, indicated by the presence of biosynthesis genes (sxt), nor psts in polish lakes has been previously studied (kobos et al., 2013). in this study, qualitative pcr methods targeting the sxta and sxtg genes and a quantitative method for the detection of sxtb, which have been previously employed in the analysis of lakes and brackish coastal waters in åland islands, finland (savela et al., 2015) were applied to investigate freshwater lakes in western poland. targeting previously studied genes allowed comparisons to existing data. to assess the potential benefit of adding a new gene target to the analysis, primers for the detection of the pst biosynthesis gene sxts were designed and used to complement existing methods. this study reports the first analysis of polish freshwater lakes for the prevalence of selected cyanobacterial paralytic shellfish toxin biosynthesis genes as well as pst prevalence and concentrations. methods study site and sampling sampling was carried out in randomly selected freshwater lakes (n=34) in western poland twice during summer 2010; first between june 6th and july 23rd (tab. 1) and again between august 20th and october 1st (tab. 2). the limnological characteristics and geographical locations of the lakes as well as the sampling procedure have been described in detail by kokociński et al. (2013). briefly, 3 to 4 subsamples were taken at 1 m vertical intervals using a 0.5 m limnos sampler starting at the depth of 1 m. one litre of the pooled subsamples from each location was transported to the laboratory and used for phytoplankton identification and biomass calculations (as described in kokociński et al., 2013) and molecular and toxin analyses. sample preparation for molecular analyses aliquots (20 ml) of the 1 l subsample were filtered onto 25 mm diameter whatman gf/c filters and stored frozen (-20°c). addition of exogenous amplification control and heat treatment (80°c, 10 min) in sterile deionized water (v=1.5 ml) water to lyse the cyanobacteria were carried out as described previously (savela et al., 2014). the resulting lysate was stored frozen (-20°c) until further analysis. spectrophotometric analysis of dna quantity and quality was not applicable to this sample type. qualitative analysis of sxta, sxtg and sxts sample lysates were analysed for the presence of pst biosynthesis genes sxta, sxtg and sxts using primers listed in tab. 3. successful amplification resulted in 648 bp, 519 bp and 382 bp pcr products, respectively. the specificity no nco mm er cia l u se on ly pst biosynthesis and genes in western polish lakes 63 of sxta and sxtg primers has been validated previously (savela et al., 2015). for this study, the in silico specificity of sxts primers was confirmed by running a query against the genbank nr database (as available in february 2011) using blastn 2.2.29+ for highly similar sequences (altschul et al., 1997) and optimized for short input sequences. the sxts target is located at bases 130-511 of the d. circinale awqc131c (dq787201) and the aphanizomenon sp. nh-5 (eu603710) sxts gene and bases 216597 of the c. raciborskii t3 (dq787200) sxts gene. the in vitro specificity was confirmed by analysing pst-producing and non-producing cyanobacterial strains following the same procedures as for sxta and sxtg (savela et al., 2015). for environmental sample analysis, sxta, sxtg and sxts-specific pcr reactions (total volume 20 µl) contained 1x phire reaction buffer and 0.4 µl phire ii hotstart dna polymerase (thermo scientific), 0.2 mm dntps (bioline), 0.5 µm appropriate forward and reverse primers and 4 µl sample lysate. one ng of d. circinale cs-537/13 genomic dna was used as the positive control. thermal cycling was performed using a bio-rad c1000 touch cycler: initial denaturation at 98°c for 30 s, then 35 cycles of 98°c 5 s, 62°c 5 s and 72°c 10 s, final extension at 72°c for 1 min (savela et al., 2015). reaction outcome was determined on 1% w/v agarose gels stained with 1x (final conc.) sybr®safe dye (thermo fisher scientific). uv illumination and gel imaging was carried out using a bio-rad geldoc xr system. tab. 1. the prevalence of sxta, sxtg, sxts, potentially pst-producing cyanobacterial species, the quantity of sxtb gene copies and the concentration and identity of psts in western polish lakes in june-july of 2010. lake sxtb sxta sxtg sxts psts a. gracile c. issatschenkoi a. flos-aquae c. raciborskii (copies l–1) (+/-) (+/-) (+/-) (µg l–1) (+/-) (+/-) (+/-) (+/-) biezdruchowo + biskupieckie biskupińskie + + bnińskie + + boczowskie 0.043 (stx) + + busko + + buszewskie 4.98×105 + + + + + bytyńskie + + długie głębokie grylewskie + + ilno + + jelonek + + + kierskie trace (stx) + kierskie małe 2.53×106 + + + trace (stx) + + kowalskie + + kruchowskie kursko + + + + lednica lubosińskie + + + lusowskie mogileńskie 8.20×104 + trace (dcstx) + + niepruszewskie niesłysz pniewskie + sarbsko 4.68×105 + + + 0.57 (stx) + + strykowskie + strzyżewskie 1.29×105 + + + + + świętokrzyskie + + szydłowskie + + tomickie + + witobelskie + żabiniec zbąszyńskie + + + no nco mm er cia l u se on ly h. savela et al.64 quantitative analysis of sxtb copy numbers the quantitative pcr (qpcr) method used to detect and measure sxtb copy numbers in cyanobacteria, and the method used to assess possible inhibition in environmental sample lysates have been described previously in detail by savela et al. (2014, 2015). primer and probe sequences are listed in tab. 3. briefly, the labelling of the sxtb detection probe with an intrinsically fluorescent terbium(iii) chelate ({2,2,´2´´, 2´´´-{{6,6´-{4´´-[2-(4-isothiocyanatophenyl)ethyl]-pyratzole-1´´,3´´-diyl}bis(pyridine)-2,2´-diyl}bis(methylenenitrilo)}-tetrakis(acetato)}tb(iii)), university of turku, turku, finland) was carried out as described by nurmi et al. (2002). the 20 µl qpcr reactions contained 0.2 u dynazyme ii hotstart polymerase, 1x dynazyme ii hs buffer, 0.2 mm dntps (finnzymes, espoo, finland), 0.4 mm both primers, 24 nm sxtb_q quencher probe, 3 nm labelled sxtb_p detection probe and 4 µl template; standard or sample lysate. quantification standards were produced from d. circinale cs-537/13 genomic dna, quantified, purified and diluted as described previously (savela et al., 2015). thermal cycling [mj research (st. bruno, canada) ptc-200 thermal cycler or bio-rad (hercules, ca, usa) c1000 touch thermal cycler] was performed as follows: initial denaturation, 95°c 5 min, then 40 cycles of 95°c 30 s, 62°c 1 min. every second cycle, from cycle 8 to cycle 40, after annealing and extab. 2. the prevalence of sxta, sxtg, sxts, potentially pst-producing cyanobacterial species, the quantity of sxtb gene copies and the concentration and identity of psts in western polish lakes in august-october of 2010. lake sxtb sxta sxtg sxts psts a. gracile c. issatschenkoi c. ovalisporum a. flos-aquae c. raciborskii (copies l–1) (+/-) (+/-) (+/-) (µg l–1) (+/-) (+/-) (+/-) (+/-) (+/-) biezdruchowo + biskupieckie biskupińskie trace (stx) + + bnińskie + + + + + boczowskie + + busko 4.31×105 + + + trace (stx) buszewskie trace (dcstx) + + bytyńskie + + długie trace (stx) głębokie grylewskie + + ilno 5.10×107 + + + 0.072 (stx) + + + jelonek + + + kierskie + + + kierskie małe + + + kowalskie 1.28×105 + + + + + kruchowskie 8.37×105 + + + trace (stx) + + kursko 1.32×105 + + + + + + + lednica lubosińskie 3.97×105 + + + + lusowskie mogileńskie 2.50×105 + + + + + + niepruszewskie niesłysz + + pniewskie 0.069 (dcstx) + + sarbsko 9.85×105 + + + + + + strykowskie + + strzyżewskie + + + świętokrzyskie + + szydłowskie 3.96×105 + + + + + tomickie 5.49×105 + + + + + + + witobelskie + + żabiniec + + + zbąszyńskie 1.14×105 + + + + + no nco mm er cia l u se on ly pst biosynthesis and genes in western polish lakes 65 tension the temperature was lowered to 35°c for 15 s, during which tb fluorescence was measured using the standard trf-tb protocol on a victor x4 2030 multilabel reader (perkinelmer, waltham, ma, usa). extraction and analysis of psts extraction of psts from water sample aliquots preserved on filters and the analysis of the resulting extracts using ion-pair high-performance liquid chromatography with post-column oxidation and fluorescence detection (hplc-fld) was carried out as described previously (savela et al., 2015). certified reference materials included c-toxins (c1-2), gonyautoxins (gtx1–4, gtx2– 3, gtx5), decarbamoyl gonyautoxins (dcgtx2-3), saxitoxin (stx), decarbamoyl saxitoxin (dcstx) and neosaxitoxin (neostx) (nrc-imb, halifax, canada). data analysis raw fluorescence measurement data from the qpcr experiments was analysed in microsoft excel 2010. amplification curves were generated by plotting well-specific signal-to-background ratios against their respective pcr cycles. background was determined for each individual reaction separately; the average signal of the first two measurements was used. cq-values were determined visually, and values deviating from the other replicates by more than one cycle were not considered in the final analysis. statistical analysis was carried out using r statistical package version 3.2.0 (r core team, 2015). results sxt genes specificities of the sxta and sxtg pcr methods as well as the properties of the sxtb qpcr assay have been described previously by savela et al. (2015). for this study, an additional primer pair targeted at the paralytic shellfish toxin biosynthetic gene sxts was applied to analysis of environmental samples. in silico analysis revealed no significant non-target hits, and no amplification was observed in cyanobacterial strains incapable of pst production (tab. 4, supplementary figure s1). the sxtb gene was amplified in 14 (41.1 %) out of the 34 lakes investigated in this study. in total, 23.5% of all samples contained the sxtb gene, detection frequency increasing towards the end of summer, from 18% in junejuly (tab. 1) to 29% in august-october (tab. 2). significant inhibition in the amplification of the exogenous amplification control was not observed. for two of these lakes, lake mogileńskie and lake sarbsko, the sxtb gene was found in samples taken both in june–july and august-–october. overall, sxtb copy numbers were low; in june-july the average was 7.4×105 copies l–1 (median 4.7×105). in august-october the average was 5.1×106 copies l–1, but the median only 4.1×105. only one relatively high sxtb copy number was measured in a sample collected at lake ilno in august 2010, 5.1×107 copies l–1. the presence of sxta, sxtg and sxts genes was confirmed in 75% (n=12) of the sxtb-positive samples. one to two target genes were absent from five samples. in four samples, either the gene sxtb or sxts was absent, once both genes sxta and sxts were not detected (tabs. 1 and 2). paralytic shellfish toxins psts were detected in a total of 13 samples. one sample, collected at lake sarbsko in july 2010 contained 0.57 µg l–1 saxitoxin. in three of the pst-positive samples concentrations between 0.043-0.072 µg l–1 were measured, and the majority (n=9) contained trace amounts of psts, the concentrations not exceeding the quantitative detection limit of the hplc-fld method (0.01 µg l–1, savela et al., 2015). only saxitoxin and decarbamoylsaxitoxin were present in the samples, other variants were not detected. psts or sxt genes were observed in 18 samples of tab. 3. primers and probes for conventional qualitative pcr and qpcr used in this study. gene name sequence (5’–3’) sxta sxta855_f* gactcggcttgttgcttcccc sxta1480_r* gccaaactcgcaacaggagaagg sxtg sxtg432_f* aatggcagatcgcaaccgctat sxtg928_r* acattcaaccctgcccattcact sxts sxts205_f° ggagtattggcgggtgactatga sxts566_r° ggtggctacttggtataactcgca sxtb sxtb_f2* tgttgtgcttgctgctctatcag sxtb_r2* cagcgttttcagcgtaycgac sxtb_p detection probe* aminoc6-caatcaaagttatgctccctatacga-phos sxtb_q quencher probe* gggagcataactttgattg-bhq®1 *savela et al., 2015; °this study. no nco mm er cia l u se on ly h. savela et al.66 the total 68. only 6 of the 68 samples contained both psts and sxt genes. no quantitative correlation between sxtb copy numbers and pst concentrations was observed (spearmans’s ρ=0.19, p=0.11). identification of potentially pst-producing cyanobacteria cyanobacterial populations in the studied lakes at the times of sampling were analysed for potentially pst-producing species (tabs. 1 and 2). particularly a. gracile was abundant, present in 75% of all samples, being as common in june-july as in august-october. the species was present in 83% and 94% of pstand sxt-gene containing samples, respectively. c. raciborskii was also commonly observed, the overall prevalence increasing from 29% to 56% towards the end of summer. it was present in 42% and 44% of pstand sxt-gene containing samples, respectively. of other potentially pst-producing species, c. issatschenkoi was observed in 26% and 29% and aphanizomenon flos-aquae ralfs ex bornet & flahault in 12% and 26% of all samples in june-july and august-october, respectively. of these two, c. issatschenkoi was present in 50% and 56% of the pstand sxt-gene containing samples. the corresponding percentages for a. flos-aquae were 8% and 22%. c. ovalisporum was present in two samples (6%) in august-october, one of which contained both psts and sxt genes. discussion in this study, potentially pst-producing cyanobacteria were for the first time indicated in polish lakes on the basis of sxt gene detection. additionally, active but lowlevel biosynthesis of psts was demonstrated by the presence of saxitoxin or decarbamoylsaxitoxin in 19% of the 68 samples tested. the results are in line with other studies investigating central european freshwaters (ballot et al., 2010; ledreux et al., 2010, jančula et al., 2014). on average, the sxt gene cluster includes 30 genes, the exact number as well as the composition and arrangement of genes within the cluster being genus-dependent (mihali et al., 2009). the gene cluster includes so-called “core”, or essential genes, without which pst biosynthesis cannot function (d’agostino et al., 2014), and the majority of these are conserved among pst-producing cyanobacterial genera (murray et al., 2011). conserved core sxt genes include sxta, sxtb, sxtg and sxts (murray et al., 2011) which were targeted in this study. these four genes are responsible for the initiation of pst biosynthesis and formation of the three central heterocycles and the second tab. 4. amplification of sxts in in toxin-producing and non-toxic cyanobacteria. species strain* sxts° toxins produced# dolichospermum circinale cs-337/01 + pst dolichospermum circinale cs-537/13 + pst dolichospermum circinale cs-530/05 nd dolichospermum circinale cs-533/12 nd anabaena cylindrica pcc 73105 nd anabaena cylindrica pcc 7938 nd dolichospermum lemmermannii var. minor niva-cya 83/1 mcyst§ dolichospermum lemmermannii var. minor niva-cya 266/1 mcyst§ cylindrospermopsis raciborskii cs-505 cyn^ cylindrospermopsis raciborskii cs-506 cyn$ cylindrospermopsis raciborskii cs-510 nd nodularia sphaerocarpa pcc 7804 l-har2-nod** nostoc sp. pcc 6310 nd nostoc sp. pcc 7422 nd microcystis aeruginosa pcc 7806 mcyst§ microcystis aeruginosa niva-cya 140 mcyst§ microcystis aeruginosa pcc 7005 nd planktothrix agardhii niva-cya 15 mcyst§ planktothrix agardhii niva-cya 299 mcyst§ planktothrix agardhii niva-cya 12 nd *strains were obtained from the australian national algae culture collection (anacc, cs), pasteur culture collection (pcc) and the norwegian institute for water research (niva). culture conditions and dna extraction have been previously described (savela et al., 2015); °this study; #data from savela et al. 2015, unless otherwise indicated; §hautala et al., 2013; ^stucken et al., 2009 and references therein; $kokociński et al., 2013; **beattie et al., 2000; nd, not detected; mcyst, microcystins; cyn, cylindrospermopsin; l-har2-nod, l-homoarginine2-nodularin. no nco mm er cia l u se on ly pst biosynthesis and genes in western polish lakes 67 guanidino group present in all pst variants known to date, and are thus necessary for pst production (kellmann et al., 2008, mihali et al., 2009, wiese et al., 2010). in the majority (75%) of sxt-positive samples (n=18) all four genes were present. the presence of all four genes suggests a functional sxt gene cluster, although other genes are still needed to complete the biosynthesis route. interestingly, in some samples (n=5) an apparent lack of one to two of the targeted sxt genes was observed. there are two possible explanations for this phenomenon. first, non-amplification might be caused by sequence dissimilarities between target and primers which could exist in natural cyanobacterial populations while being absent from the sxt gene sequences available at sequence databases. both the sxta and sxtg primer sequences are identical with their respective binding sites in aphanizomenon sp. and c. raciborskii sequences, but nevertheless sxta failed to amplify in one sample despite other sxt genes being present, indicating critical dissimilarities between the primers and the target. both sxtb and sxts forward primers share at most one known internal nucleotide mismatch with the corresponding sequences in the abovementioned species. an internal mismatch can affect reaction efficiency, but is not sufficient to abolish amplification, as shown by the robust amplification of sxts in d. circinale, which shares an identical primer binding site, including the one mismatch, with aphanizomenon sp. nh-5. second, mutations in a larger scale, including deletions and insertions within the sxt gene cluster would result in a negative outcome in pcr analysis. such mutations would also result in an inactive gene cluster and effectively a non-toxic cyanobacterial strain. other cyanotoxin genes, most notably the microcystin biosynthesis genes are known to occasionally contain mutations rendering the biosynthesis pathway inactive (christiansen et al., 2006; noguchi et al., 2009; ostermaier and kurmayer, 2009). this suggests that similar mutations might be found also in pst-producing cyanobacteria. indeed, previous studies have shown that the presence of sxt genes e.g. sxta, does not always indicate detectable pst production in cyanobacteria; sxta has been detected in nonpst-producing aphanizomenon, anabaena and anabaenopsis strains (ballot et al., 2010; casero et al., 2014). similarly, in this study none of the analysed samples in which a partial set of sxt genes was observed contained psts. however, the general lack of correlation between sxt gene presence and quantity and pst concentrations means that definitive conclusions on the presence of mutated sxt gene clusters cannot be drawn, and to confirm the hypothesis a suspected mutant strain would have to be isolated and the sxt genes characterized in detail. an overall lack of correlation between sxt gene presence and copy number quantity and pst concentrations in the investigated lakes was apparent. only a few surveys of environmental samples without prior strain isolation have been carried out, however, findings similar to this study have been obtained in greece, where amplification of the sxti gene was twice as common as pst detection (gkelis and zaoutsos, 2014) and during a three-year survey of lakes in åland islands in finland using the same methods as in this study (savela et al., 2015) with the exception of the new sxts primers. the observed lack of correlation is not limited to europe; in australia, neither the presence of sxta nor the gene’s increasing copy numbers did indicate a detectable presence or growing concentrations of psts in the environment (al-tebrineh et al., 2012; bowling et al., 2013). comparisons of qpcr and hplc-fld data are complicated by the fact that qpcr can be extremely sensitive (wittwer et al., 1997), whereas psts can be difficult to analyse in complex environmental sample extracts. for this reason, molecular analysis of pst biosynthesis genes have been proposed as an easier alternative for screening environmental samples for potential pst production. in this study, in addition to using hplc-fld, samples were originally analysed for psts using a hydrophilic interaction liquid chromatographytandem mass spectrometry (hilic-ms/ms) method (diener et al., 2007; savela et al., 2015). due to the very low pst concentrations, and similarly to previous work reported by savela et al. (2015), only the hplc-fld method could be used to detect the toxins; its sensitivity was far superior compared to that of the hilic-ms/ms method, which employed an ion-trap ms instrument. notwithstanding methodological limitations, the findings from this and previous studies suggest regulation of pst production at the transcriptional and/or translational level, rather than toxin biosynthesis being a direct consequence of the biosynthetic capability encoded in the genome. environmental factors including osmotic pressure, ionic stress and availability of organic nitrogen have been shown to affect sxt gene expression, pst production and export in aphanizomenon spp., d. circinale and c. raciborskii (dias et al., 2002; pomati et al., 2004; casero et al., 2014; ongley et al., 2016), suggesting complex regulation mechanisms of toxin biosynthesis tied to the processes of primary metabolism. therefore the simple approach of measuring sxt gene presence or copy number would not serve as a reliable indicator of pst production; measuring gene expression instead might predict active pst biosynthesis better. of the potentially pst-producing cyanobacteria, a. gracile was present in 83% and 94% of sxt geneand/or pst-containing samples, respectively. the role of a. gracile in pst production is supported by its confirmed status as a pst-producing species in e.g. north-eastern germany (ballot et al., 2010). previously, a positive correlation between a. gracile biomass and cylindrospermopsin, another cyanobacterial toxin, has already been no nco mm er cia l u se on ly h. savela et al.68 shown (kokociński et al., 2013). these observations suggest that a. gracile populations in western polish lakes may contain both cylindrospermopsinand pst-producing strains. however, the involvement of other aphanizomenon species, which often co-occurred with a. gracile in the lakes included in this study, in pst production cannot be ruled out. further study into the aphanizomenon spp. described here would be required to definitely confirm the identity and possible role in pst production in polish lakes. the second most common species to cooccur with sxt genes and/or psts was c. raciborskii. to date, the species has not been associated with pst production in european freshwaters, and conclusions of its possible role as a pst producer in western polish lakes cannot be made without further investigation. conclusions in this study, both the genetic capability for pst biosynthesis and active pst production in selected freshwater lakes in western poland was demonstrated for the first time. the most likely potentially pst-producing species is a. gracile, however, further studies on isolated strains need to be conducted to confirm this hypothesis. the low sxtb gene copy numbers and pst concentrations indicate that pst producers constituted a relatively minor subpopulation in the studied waterbodies. the non-amplification of certain sxt genes in the presence of others suggested the possible presence of inactive sxt gene clusters in the studied cyanobacterial populations. since no significant correlation between genes sxta, sxtb, sxtg and sxts, and measured psts was observed, it was concluded that analysis of sxt genes indicates the presence of potentially pst-producing cyanobacteria, but is not suitable for predicting active pst biosynthesis in the studied polish lakes. acknowledgments this work was financially supported by the finnish funding agency for innovation (tekes) grant 40013/10. the authors gratefully acknowledge the european cooperation in science and technology, cost action es 1105 ‘‘cyanocost-cyanobacterial blooms and toxins in water resources: occurrence, impacts and management’’ for adding value to this study through networking and knowledge sharing with european experts and researchers in the field. references al-tebrineh j, merrick c, ryan d, humpage a, bowling l, neilan ba, 2012. community composition, toxigenicity, and environmental conditions during a cyanobacterial bloom occurring along 1,100 kilometers of the murray river. appl. environ. microbiol. 78:263-272. altschul sf, madden tl, schaffer aa, zhang j, zhang z, miller w, lipman dj, 1997. gapped blast and psi-blast: a new generation of protein database search programs. nucleic acids res. 25:3389-3402. ballot a, fastner j, wiedner c, 2010. paralytic shellfish poisoning toxin-producing cyanobacterium aphanizomenon gracile in northeast germany. appl. environ. microbiol. 76:1173-1180. beattie k, kaya k, codd g, 2000. the cyanobacterium nodularia pcc 7804, of freshwater origin, produces [lhar2]nodularin. phytochemistry 54:57-61. borges hlf, branco lhz, martins md, lima cs, barbosa pt, lira gast, bittencourt-oliveira, mc, molica rjr, 2015. cyanotoxin production and phylogeny of benthic cyanobacterial strains isolated from the northeast of brazil. harmful algae 43:46-57. bowling lc, merrick c, swann j, green d, smith g, neilan ba, 2013. effects of hydrology and river management on the distribution, abundance and persistence of cyanobacterial blooms in the murray river, australia. harmful algae 30:27-36. carmichael ww, evans wr, yin qq, bell p, moczydlowski e, 1997. evidence for paralytic shellfish poisons in the freshwater cyanobacterium lyngbya wollei (farlow ex gomont) comb. nov. appl. environ. microbiol. 63:3104-3110. casero mc, ballot a, agha r, quesada a, cirés s, 2014. characterization of saxitoxin production and release and phylogeny of sxt genes in paralytic shellfish poisoning toxin-producing aphanizomenon gracile. harmful algae 37:28-37. christiansen g, kurmayer r, liu q, börner t, 2006. transposons inactivate biosynthesis of the nonribosomal peptide microcystin in naturally occurring planktothrix spp. appl. environ. microbiol. 72:117-123. cirés s, wörmer l, ballot a, agha r, wiedner c, velazquez d, casero mc, quesada a, 2014. phylogeography of cylindrospermopsin and paralytic shellfish toxin-producing nostocales cyanobacteria from mediterranean europe (spain). appl. environ. microbiol. 80:1359-1370. cirés s, ballot a, 2016. a review of the phylogeny, ecology and toxin production of bloom-forming aphanizomenon spp. and related species within the nostocales (cyanobacteria). harmful algae 54:21-43. d’agostino pm, moffitt mc, neilan ba, 2014. current knowledge of paralytic shellfish toxin biosynthesis, molecular detection and evolution, p. 251-280. in: rossini, g.p. (ed.), toxins and biologically active compounds from microalgae.1. crc press. dias e, pereira p, franca s, 2002. production of paralytic shellfish toxins by aphanizomenon sp. lmecya 31 (cyanobacteria). j. phycol. 38:705-712. diener m, erler k, christian b, luckas b, 2007. application of a new zwitterionic hydrophilic interaction chromatography column for determination of paralytic shellfish poisoning toxins. j. sep. sci. 30:1821-1826. gkelis s, zaoutsos n, 2014. cyanotoxin occurrence and potentially toxin producing cyanobacteria in freshwaters of greece: a multi-disciplinary approach. toxicon 78:1-9. no nco mm er cia l u se on ly pst biosynthesis and genes in western polish lakes 69 hautala h, lamminmäki u, spoof l, nybom s, meriluoto j, vehniäinen m, 2013. quantitative pcr detection and improved sample preparation of microcystin-producing anabaena, microcystis and planktothrix. ecotoxicol. environ. saf. 87:49-56. jančula d, straková l, sadílek j, maršálek b, babica p, 2014. survey of cyanobacterial toxins in czech water reservoirs the first observation of neurotoxic saxitoxins. environ. sci. pollut. r. 21:8006-8015. kaas h, henriksen p, 2000. saxitoxins (psp toxins) in danish lakes. water res. 34:2089-2097. kao cy, 1986. structure-activity relations of tetrodotoxin, saxitoxin, and analogues ann. n. y. acad. sci. 479:52-67. kellmann r, mihali tk, jeon yj, pickford r, pomati f, neilan ba, 2008. biosynthetic intermediate analysis and functional homology reveal a saxitoxin gene cluster in cyanobacteria. appl. environ. microbiol. 74:4044-4053. kobos j, błaszczyk a, hohlfeld n, toruńska-sitarz a, krakowiak a, hebel a, sutryk k, grabowska m, toporowska m, kokociński m, messyasz b, rybak a, napiórkowska-krzebietke a, nawrocka l, pełechata a, budzyńska a, zagajewski p, mazur-marzec h, 2013. cyanobacteria and cyanotoxins in polish freshwater bodies. oceanol. hydrobiol. stud. 42:358-378. kokociński m, soininen j, 2012. environmental factors related to the occurrence of cylindrospermopsis raciborskii (nostocales, cyanophyta) at the north-eastern limit of its geographical range. eur. j. phycol. 47:12-21. kokociński m, mankiewicz-boczek j, jurczak t, spoof l, meriluoto j, rejmonczyk e, hautala h, vehniäinen m, pawełczyk j, soininen j, 2013. aphanizomenon gracile (nostocales), a cylindrospermopsin-producing cyanobacterium in polish lakes. environ. sci. pollut. res. 20:52435264. lagos n, onodera h, zagatto pa, andrinolo d, azevedo sm, oshima y, 1999. the first evidence of paralytic shellfish toxins in the fresh water cyanobacterium cylindrospermopsis raciborskii, isolated from brazil. toxicon 37:1359-1373. ledreux a, thomazeau s, catherine a, duval c, yéprémian c, marie a, bernard c, 2010. evidence for saxitoxins production by the cyanobacterium aphanizomenon gracile in a french recreational water body. harmful algae 10:88-97. li r, carmichael ww, pereira p, 2003. morphological and 16s rrna gene evidence for reclassification of the paralytic shellfish toxin producing aphanizomenon flos-aquae lmecya 31 as aphanizomenon issatschenkoi (cyanophyceae). j. phycol. 39:814-818. llewellyn le, 2006. saxitoxin, a toxic marine natural product that targets a multitude of receptors. nat. prod. rep. 23:200222. mankiewicz-boczek j, izydorczyk k, romanowska-duda z, jurczak t, stefaniak k, kokociński m, 2006. detection and monitoring toxigenicity of cyanobacteria by application of molecular methods. environ. toxicol. 21:380-387. mihali tk, kellmann r, neilan ba, 2009. characterisation of the paralytic shellfish toxin biosynthesis gene clusters in anabaena circinalis awqc131c and aphanizomenon sp. nh5. bmc biochem. 10:8. murray sa, mihali tk, neilan ba, 2011. extraordinary conservation, gene loss, and positive selection in the evolution of an ancient neurotoxin. mol. biol. evol. 28:1173-1182. noguchi t, shinohara a, nishizawa a, asayama m, nakano t, hasegawa m, harada k, nishizawa t, shirai m, 2009. genetic analysis of the microcystin biosynthesis gene cluster in microcystis strains from four bodies of eutrophic water in japan. j. gen. appl. microbiol. 55:111-123. nurmi j, wikman t, karp m, lövgren t, 2002. high-performance real-time quantitative rt-pcr using lanthanide probes and a dual-temperature hybridization assay. anal. chem. 74:3525-3532. ongley se, pengelly jjl, neilan ba, 2016. elevated na+ and ph influence the production and transport of saxitoxin in the cyanobacteria anabaena circinalis awqc131c and cylindrospermopsis raciborskii t3. environ. microbiol. 18: 427-438. ostermaier v, kurmayer r, 2009. distribution and abundance of nontoxic mutants of cyanobacteria in lakes of the alps. microb. ecol. 58:323-333. pereira p, onodera h, andrinolo d, franca s, araujo f, lagos n, oshima y, 2000. paralytic shellfish toxins in the freshwater cyanobacterium aphanizomenon flos-aquae, isolated from montargil reservoir, portugal. toxicon 38:1689-1702. pereira p, li r, carmichael w, dias e, franca s, 2004. taxonomy and production of paralytic shellfish toxins by the freshwater cyanobacterium aphanizomenon gracile lmecya40. eur. j. phycol. 39:361-368. pomati f, rossetti c, manarolla g, burns bp, neilan ba, 2004. interactions between intracellular na+ levels and saxitoxin production in cylindrospermopsis raciborskii t3. microbiology 150:455-461. r core team, 2015. r: a language and environment for statistical computing, http://www.r-project.org. r foundation for statistical computing, vienna, austria. rajaniemi p, komárek j, willame r, hrouzek p, kaštovská k, hoffmann l, sivonen k, 2005. taxonomic consequences from the combined molecular and phenotype evaluation of selected anabaena and aphanizomenon strains. algological studies 117:371-391. rapala j, robertson a, negri ap, berg ka, tuomi p, lyra c, erkomaa k, lahti k, hoppu k, lepistö l, 2005. first report of saxitoxin in finnish lakes and possible associated effects on human health. environ. toxicol. 20:331-340. savela h, spoof l, perälä n, preede m, lamminmäki u, nybom s, häggqvist k, meriluoto j, vehniäinen m, 2015. detection of cyanobacterial sxt genes and paralytic shellfish toxins in freshwater lakes and brackish waters on åland islands, finland. harmful algae 46:1-10. savela h, vehniäinen m, spoof l, nybom s, meriluoto j, lamminmäki u, 2014. rapid quantification of mcyb copy numbers on dry chemistry pcr chips and predictability of microcystin concentrations in freshwater environments. harmful algae 39:280-286. smith fm, wood sa, van ginkel r, broady pa, gaw s, 2011. first report of saxitoxin production by a species of the freshwater benthic cyanobacterium, scytonema agardh. toxicon 57:566-573. stucken k, murillo aa, soto-liebe k, fuentes-valdes jj, mendez ma, vasquez m, 2009. toxicity phenotype does not correlate with phylogeny of cylindrospermopsis raciborskii strains. syst. appl. microbiol. 32:37-48. no nco mm er cia l u se on ly h. savela et al.70 su z, sheets m, ishida h, li f, barry wh, 2004. saxitoxin blocks l-type ica. j. pharmacol. exp. ther. 308:324-329. wacklin p, hoffmann l, komarek j, 2009. nomenclatural validation of the genetically revised cyanobacterial genus dolichospermum (ralfs ex bornet et flahault) comb. nova. fottea 9:59-64. wang j, salata jj, bennett pb, 2003. saxitoxin is a gating modifier of herg k+ channels. j. gen. physiol. 121:583-598. wiese m, d’agostino pm, mihali tk, moffitt mc, neilan ba, 2010. neurotoxic alkaloids: saxitoxin and its analogs mar. drugs 8:2185-2211. wittwer ct, herrmann mg, moss aa, rasmussen rp, 1997. continuous fluorescence monitoring of rapid cycle dna amplification. biotechniques 22:130-1, 134-8. wörmer l, cirés s, agha r, verdugo m, de hoyos c, quesada a, 2011. first detection of cyanobacterial psp (paralytic shellfish poisoning) toxins in spanish freshwaters. toxicon 57:918-921. yunes js, de la rocha s, giroldo d, silveira sbd, comin r, bicho mds, melcher ss, sant’anna cl, vieira aah, 2009. release of carbohydrates and proteins by a subtropical strain of raphidiopsis brookii (cyanobacteria) able to produce saxitoxin at three nitrate concentration. j. phycol. 45:585-591. zapomělová e, skácelová o, pumann p, kopp r, janeček e, 2012. biogeographically interesting planktonic nostocales (cyanobacteria) in the czech republic and their polyphasic evaluation resulting in taxonomic revisions of anabaena bergii ostenfeld 1908 (chrysosporum gen. nov.) and a. tenericaulis nygaard 1949 (dolichospermum tenericaule comb. nova). hydrobiologia 698:353-365. no nco mm er cia l u se on ly layout 1 advances in oceanography and limnology, 2015; 6(1/2): 76-77 book review doi: 10.4081/aiol.2015.5479 to usher in a shift, the new editorial shape and the new publisher of advances in limnology and oceanography, we could not have hoped better than providing our members and authors a review of a special issue of philosophical transactions of the royal society b devoted to the regime shifts in the marine environment. when, in the mid-80s of the last century, the scientific community has begun to alert the world’s governments about the imminence and the possible consequences of global change on the worldwide oceans, the underwater landscape was indeed not too different from what, in the late 50s, the first and pioneering underwater filming had shown. today, little more than 30 years after those first aware signs of alarm, we are faced with the blatant evidence of a change of the underwater seascape (boström et al., 2011) that not only hit the coastal seas that we can all look through the glass of a mask, but has now begun to affect manifestly even the deep ocean (danovaro et al., 2013). we also know that such changes were not always smooth, but also sudden, tangible, measurable and dramatic in their manifestation (scheffer and carpenter, 2003). these sudden changes may affect populations, communities or whole ecosystems and are known as regime shifts (scheffer, 2009). the january 2015 special issue of philosophical transactions of the royal society b entitled marine regime shifts around the globe: theory, drivers and impacts edited by alessandra conversi, christian möllmann, carl folke and martin edwards, includes 14 articles introduced by a foreword perspective (möllman et al., 2015) and a closing executive summary with holistic ambitions (conversi et al., 2015) that collects the summa of knowledge, uncertainties and gaps accumulated to date about the understanding of regime shifts in the marine environment, their causes and the harmful consequences they can have on marine ecosystems and related services. overall, the special issue is cleverly organized according to a logical flow that highlights, first of all, the theoretical aspects and the emerging issues of the more recent research on marine regime shifts, identifies their main drivers, and provides a list of tools to adaptively manage ecosystems affected by regime shifts; furthermore, it also provides a critical revision of the ‘regime shift’ concept, as well as insights and ideas for future directions of investigation. prestigious worldwide-recognized marine ecologists, including the most known experts of marine regime shifts, signed critical reviews dedicated to indicators of resilience (dakos et al., 2015), mechanisms underlying the formation of alternative stable states (gårdmark et al. 2015), the theoretical bases for the prediction of regime shifts led by climate (beaugrand, 2015), the different types of regime shifts and their causes (jouffray et al., 2015), the resilience of coral reefs to global warming (bozec et al., 2015), the role of overgrazing by sea urchins (ling et al., 2015) and trophic cascades (pershing et al., 2015), the planetary value of regime shifts in the ocean (beaugrand et al., 2015), the issues of scale of regime shifts drivers (fisher et al., 2015) and their impacts on ecosystem services (rocha et al., 2015). the following four articles are devoted to the management of regime shifts and, in particular, to the opportunities of an ecosystem-based management approach (levin and möllmann, 2015), to the major issues associated with fisheries management (king et al., 2015), to the identification of appropriate thresholds of management (kelly et al., 2015), to the issues related to the fishing effort by the soviet union (österblom and folke, 2015). the final chapter, in an attempt to provide a holistic approach, sums up the known and unknown about regime shifts, their triggering mechanisms, their development, and their ecological consequences, ultimately proposing the future research challenges for the coming years. not to forget that the special issue is dedicated to frédéric ibañez esteemed numerical ecologist of the laboratory of oceanography of villefranche-sur-mer. in just over 150 pages this special issue of philosophical transactions b edited by conversi et al. represents a seminal reference not only for those who will decide to go into the study and understanding of regime shifts in the marine environment, but also a very useful compendium for students in advanced phd courses in discieverything you wanted to know about regime shifts in the marine environment and have not dared to ask, yet antonio pusceddu,1 silvia bianchelli2 1dipartimento di scienze della vita e dell’ambiente, università degli studi di cagliari, via fiorelli, 1, 09126 cagliari, italy; 2dipartimento di scienze della vita e dell’ambiente, università politecnica delle marche, via brecce bianche, 60131 ancona, italy no nco mm er cia l u se on ly 77 a. pusceddu and s. bianchelli plines such as biological oceanography or marine ecology. ultimately, a must read for any marine ecologist. references beaugrand g, 2015. theoretical basis for predicting climate-induced abrupt shifts in the oceans. phil. trans. r. soc. b 370: 20130264. beaugrand g, conversi a, chiba s, edwards m, fonda-umani s, greene c, mantua n, otto sa, reid pc, stachura mm, stemmann l, sugisaki h, 2015. synchronous marine pelagic regime shifts in the northern hemisphere. phil. trans. r. soc. b 370:20130272. bostrom̈ c, pittman sj, simenstad c, kneib rt, 2011. seascape ecology of coastal biogenic habitats: advances, gaps, and challenges. mar. ecol. progr. ser. 427:191–217. bozec y-m, mumby pj, 2015. synergistic impacts of global warming on the resilience of coral reefs. phil. trans. r. soc. b 370:20130267. conversi a, dakos v, gårdmark a, ling s, folke c, mumby pj, greene c, edwards m, blenckner t, casini m, pershing a, möllmann a, 2015. a holistic view of marine regime shifts. phil. trans. r. soc. b 370:20130279. dakos v, carpenter sr, van nes eh, scheffer m, 2015. resilience indicators: prospects and limitations for early warnings of regime shifts. phil. trans. r. soc. b 370:20130263. danovaro r, snelgrove p, tyler p, 2013. challenging the paradigms of deep-sea ecology. trends ecol. evol. 29:465-475. fisher jad, casini m, frank kt, möllmann c, leggett wc, daskalov g, 2015. the importance of within-system spatial variation in drivers of marine ecosystem regime shifts. phil. trans. r. soc. b 2015 370:20130271. gårdmark a, casini m, huss m, van leeuwen a, hjelm j, persson l, de roos am, 2015. regime shifts in exploited marine food webs: detecting mechanisms underlying alternative stable states using size-structured community dynamics theory. phil. trans. r. soc. b 370:20130262. jouffray j-b, nyström m, norström av, williams id, wedding lm, kittinger jn, williams gj, 2015. identifying multiple coral reef regimes and their drivers across the hawaiian archipelago. phil. trans. r. soc. b 370:20130268. kelly rp, erickson al, mease la, battista w, kittinger jn, fujita r, 2015. embracing thresholds for better environmental management. phil. trans. r. soc. b 370:20130276. king jr, mcfarlane ga, punt ae, 2015. shifts in fisheries management: adapting to regime shifts. phil. trans. r. soc. b 370:20130277. levin ps, möllmann c, 2015. marine ecosystem regime shifts: challenges and opportunities for ecosystem-based management. phil. trans. r. soc. b 370:20130275. ling sd, scheibling re, rassweiler a, johnson cr, shears n, connell sd, salomon ak, norderhaug km, pérez-matus a, hernández jc, clemente s, blamey lk, hereu b, ballesteros e, sala e, garrabou j, cebrian e, zabala m, fujita d, johnson le, 2015. global regime shift dynamics of catastrophic sea urchin overgrazing. phil. trans. r. soc. b 370: 20130269. möllmann c, folke c, edwards m, conversi a, 2015. introduction: marine regime shifts around the globe: theory, drivers and impacts. phil. trans. r. soc. b 2015 370:20130260. österblom h, folke c, 2015. globalization, marine regime shifts and the soviet union. phil. trans. r. soc. b 2015 370: 20130278. pershing aj, mills ke, record nr, stamieszkin k, wurtzell kv, byron cj, fitzpatrick d, golet wj, koob e, 2015. evaluating trophic cascades as drivers of regime shifts in different ocean ecosystems. phil. trans. r. soc. b 370:20130265. rocha j, yletyinen j, biggs r, blenckner t, peterson g, 2015. marine regime shifts: drivers and impacts on ecosystems services. phil. trans. r. soc. b 370:20130273. scheffer m, 2009. critical transitions in nature and society. princeton university press, princeton, nj, usa: 400 pp. scheffer m, carpenter sr, 2003. catastrophic regime shifts in ecosystems: linking theory to observation. trends ecol. evol. 18:648-656. no nco mm er cia l u se on ly layout 1 introduction analysis of subfossil cladocera (crustacea: branchiopoda) is widely used in paleolimnology given its potential to reconstruct past environmental conditions (korhola and rautio, 2001; nevalainen and rautio, 2014; zawiska et al., 2015). cladocerans are used as indicators of several abiotic and biotic environmental variables (rumes et al., 2011; chen et al., 2014), as they are very sensitive to changes in total phosphorus concentrations (amsinck et al., 2005; chen et al., 2010), water depth (korhola et al., 2005; nevalainen et al., 2011; gałka et al., 2014), temperature (lotter et al., 1997; korhola, 1999; mirosław-grabowska and zawisza, 2013; nevalainen et al., 2013; zawiska et al., 2015), ph (locke and sprules, 2000; zawiska et al., 2013). the crucial step in subfossil cladocera analysis is the correct taxonomical identification of the remaisns at the species level, which is usually based on the use of light microscope (magnification 100-400x) and several determination keys ( alonso, 1996; szeroczyńska and sarmajakorjonen, 2007; korosi and smol, 2012). as the light microscopy allows to observe the remains in two dimensions, the body sculpture appears to be only a pattern on the surface of the carapace. the microstructural characteristics of the chitinous remains can be observed in threedimensional appearance only by scanning electron microscope (sem), which enables to create images by scanning the surface with a focused beam of electrons (goldstein et al., 2003). the magnifications obtained with sem are much greater than those of light microscopy and reach 100,000x. sem is commonly used in taxonomy of living cladocera to describe morphological features such as the limb setae, the lateral pores or the shell denticles (sinev et al., 2005; sinev and elmoor-loureiro, 2010). cladocera for sem observations are usually collected from water by using a plankton net and dried using either a wide range of alcohol percentages (70%, 90%, 95%, 100%) (duigan, 1992; nandini et al., 2009), or the strong reagent hexamethyldisilazane (laforsch and tollrian, 2000; sousa et al., 2015; juračka et al., 2016). saha et al. (2011) recently presented a new simplified procedure, where specimens collected from water samples are washed in distilled water, dried in the room temperature for 30 mins, coated with gold palladium and examined with sem. sem images are also frequently used in paleolimnology to study sediment properties, such as the origin of carbonates (terrestrial or autogenic), porosity, composition of lamination and microfossil taxa identification (kemp et al., 2001; martín-serrano et al., 2009; wetzel, 2013; kirillova et al., 2016). they are also very useful in observing small morphological details of microorganism and advances in oceanography and limnology, 2016; 7(2): 177-183 article doi: 10.4081/aiol.2016.6218 this work is licensed under a creative commons attribution-noncommercial 4.0 international license (cc by-nc 4.0). exploring the world of micro sculptures subfossil cladocera remains under the sem izabela zawiska,1* edyta zawisza,2 marta wojewódka,2 artem y. sinev3 1department of geoecology and climatology, institute of geography and spatial organization, polish academy of sciences, twarda 51/55, pl-00-818 warsaw, poland; 2institute of geological sciences, polish academy of sciences, twarda 51/55, pl-00-818 warsaw, poland; 3department of invertebrate zoology, biological faculty, lomonosov moscow state university, leninskie gory, 119991 moscow, russia *corresponding author: izawiska@twarda.pan.pl abstract the scanning electron microscope (sem) is widely used for the identification of microstructural characteristics and morphology of different microorganisms. common procedures are based and developed for remains of living species. this paper presents an effective method for drying and preparing subfossil cladocera remains for sem observation, which has been recently adapted and tested on several samples originating from different american and european lakes. this method results to be fast and cheap, as it excludes the use of expensive and toxic reagents. moreover, it allows to recognize the micro sculpture and other species specific characteristics present on the different body parts of the cladocera remains. the present contribution provides 29 high quality pictures of 12 cladoceran species at magnification between 200x and 11,000x. sem images reveal that the patterns observed on the shells under the light microscope actually are always three dimensional structures. key words: sem; subfossil cladocera; micro sculpture; chitin. received: august 2016. accepted: december 2016. no n c om me rci al us e o nly 178 i. zawiska et al. often help to identify remains such as diatoms (battarbee, 2001) or testate amoebae (beyens and meisterfeld, 2001). therefore, the methodology for preparing the subfossil remains of these organism for sem observation is well established (jiang et al., 2015). on the contrary, cladocera subfossil remains are still rarely observed under the sem (kirillova et al., 2016). in fact, although cladocera skeleton is composed of fairly hard chitin, the typical thickness variation depending on the species, body part and lake environment conditions, make the cladocera preparation for sem more complicated and time-consuming (andrademorraye et al., 2004). the sediment samples should be firstly prepared according to standard procedure (frey, 1986), then remains have to be picked up from the samples and washed several time with distilled water. after that remains should be put to osmium tetroxide for 2 h, washed in distilled water again and submitted to the ethanol dehydration sequence (andrade-morraye et al., 2004). when the remains are acquired from unconsolidated sediments they have to be submitted to dehydration in graded alcohols solutions (kirillova et al., 2016). in our research we aimed at testing whether the simplified method proposed by saha et al. (2011) for aquatic samples could be applied also to subfossil cladocera remains in order to obtain good quality pictures. methods subfossil cladocera remains for sem observation were obtained from sediment samples using two approaches. in the first one fresh sediment from different lakes located in central and south america was analysed (i.e. from lake comendador, lake chicabal, lake quexil (guatemala), lake emiliano zapata (yucatan peninsula, mexico), lake los negritos, lake verde (salvador), lake madre vieja, (honduras), lake san martin, (argentina). remains were picked directly from the unconsolidated upper first cm of surface sediments (1 cm3), diluted with distilled water and put into a petri dish. cladocera remains were pick out using a pipette (in the drop of water) under the dissecting microscope and directly put on the sem microscope stubs covered with a carbon adhesive tape. in the second approach the sediment samples for sem observation were obtained from european lakes, i.e. atnsjøen (norway), czechowskie (poland) and suchar iv (poland). samples were taken from sediment cores and chemically prepared for subfossil cladocera analysis according to standard procedures (szeroczyńska and sarmaja-korjonen, 2007). the amount 1 cm3 of fresh sediment was treated with hot 10% koh for 20 min using a magnetic stirrer in order to deflocculate the material and remove humic substances. thereafter the carbonates were removed using 10% hcl. the remaining material was sieved through 33 µm mesh and diluted in 10 cm3 distilled water. the subfossil remains were removed consecutively with a pipette from the cleaned sediment and directly put on the sem microscope stubs covered with a carbon adhesive tape. the remains obtained from both superficial fresh sediments and cleaned core material were left to dry at the room temperature for 48 hs. when dried they were put into the sputter coater sc7620 for 120 s and coated with a gold-palladium. the sample coating with an electric conducting material is necessary in order to avoid the accumulation of electrostatic charge at the sample surface (sinev et al., 2005; sinev and elmoor-loureiro, 2010). after the specimens were coated they were put into the scanning electron microscope (jeol jsm-6610lv) chamber and observed in high vacuum, using sei mode, voltage 20 kv. results figures 1-8 show 29 good quality images of subfossil cladocera remains belonging to 12 cladocera species. the pictures were taken at magnification ranging from 200x to 11,000x. the remains from both fresh sediments and from the sediment cores have different state of preservation, independently from the age of the sample. the remains of chydorus spp. (leach) preserved well and therefore easier to be photographed, compared to the other observed cladocera species (figs. 1 and 2). the sem pictures allowed to observe magnificent sculpture of ceriodaphnia spp. (dana) and simpocephallus ephippia (schoedler) (figs. 3 and 4). the delicate structure of alonella excisa (fischer) shell and leydigiopsis ornata (daday) (fig. 5), the deep carvings of graptoleberis testudinaria (fischer) and monospilus dispar (sars) (fig. 6), as well as characteristic triangle on the paralona pigra (sars) shell are well documented (fig. 1). the sem pictures revealed the three-dimensional aspect of the head pores of alona ossiani (sinev), bosmina (e.) coregoni (baird) and bosmina (e.) longispina (leydig) (fig. 7). on the contrary, the specimens form lake suchar iv showed high level of degradation, and diminished sculpture of the remains (fig. 8). this might be possibly due to the fact that these sediment samples were prepared for subfossil cladocera analysis already five years ago. discussion the simplified procedure of preparing specimens for sem observation proposed by saha et al. (2011) was tested on different remains of cladocera species from several american and european lakes. this procedure is based on the concept developed for the remains of living species (sinev et al., 2005; van damme and dumont, no n c om me rci al us e o nly exploring the world of micro sculptures subfossil cladocera remains under the sem 179 fig. 1. sem images of cladocera remains. a) chydorus sphericus shell, magnification 270x (lake comendador, guatemala). b) chydorus sphericus shell sculpture, magnification 4000x (lake comendador, guatemala). c) paralona pigra shell, magnification 330x (lake san martin, argentina). d) paralona pigra shell, triangular anterior accessory flange on the anteriror-ventral margin, magnification 950x (lake san martin, argentina). e) chydorus spp., magnification 500x (lake comendador, guatemala). f) chydorus spp., magnification 270x (lake comendador, guatemala). fig. 2. sem images of cladocera remains of ceriodaphnia spp. ephippium (lake emiliano zapata, yucatan peninsula, mexico). magnification: a) 160x; b) 650x; c) 900x; d) 4000x. no n c om me rci al us e o nly 180 i. zawiska et al. 2007; sinev and elmoor-loureiro, 2010; kotov, 2013; sousa et al., 2015). the time for drying the remains suggested by saha et al. (2011) was not long enough in the case of fragmented parts of the cladoceran body found in the studied samples. since most of them were very small and difficult to pick out from the sediment sample with a needle, they were put on the stage with a pipette, in a fairly large drop of water. therefore the prolongation of the drying time was necessary. from all types of examined cladocera remains, ephippia showed the best reservation of structure and ornamentation, as their chitinous envelope is thick and less prone for mechanic destruction. it was also noted that not all subfossil cladocera remains were suitable for sem observation, as some were so thin that the specimens were barely visible. in addition, it was recognized that the subfossil material for sem observation should be pick out from the freshly prepared sediment sample, as the remains slowly degrade and the chitin structure become less prominent after sediment preparation (fig. 8). the sem images clearly showed that the patterns on the shells observed under the light microscope always correspond to three dimensional structures. conclusions the simplified method of preparing subfossil cladocera was applied on different samples from several lakes. the presented method resulted to be simple, cheap and allowed to create high quality images of all types of refig. 3. sem images of cladocera remains of simpocephalus spp. ephippium (lake chicabal, guatemala). magnification: a) 150x; b) 950x; c) 3500x. fig. 4. sem images of cladocera remains. a) alonella excisa shell, magnification 400x (lake quexil, guatemala). b) alonella excisa shell sculpture, magnification 1300x (lake quexil, guatemala). c) leidigiopsis ornata head, magnification 230x (lake los negritos, salvador). d) leidigiopsis ornata head sculpture, magnification 1600x (lake los negritos, salvador). no n c om me rci al us e o nly exploring the world of micro sculptures subfossil cladocera remains under the sem 181 mains even in fairly high magnifications up to 11,000x. although the procedure developed for the remains of living species revealed to be effective also for subfossil cladocera, it appeared necessary to prolong the drying time when working with subfossil remains. moreover, the samples should be prepared just before the sem analysis, in order to prevent the degradation of the micro sculpture after the cleaning procedure. fig. 5. sem images of cladocera remains. a) graptoleberis testudinaria shell, magnification 300x (lake verde, salvador). b) graptoleberis testudinaria head, magnification 370x (lake verde, salvador). c) monospilus dispar head, magnification 400x (czechowskie lake, poland). fig. 6. sem images of cladocera remains of alona ossiani head (lake madre vieja, honduras). magnification: a) 190x; b) 1600x; c,d) 11,000x. fig. 7. sem images of cladocera remains from lake atnsjøen, norway. a) bosmina e.coregoni head, magnification 500x. b) bosmina e.coregoni head pore, magnification 3300x. c) bosmina e. longispina head, magnification 800x. fig. 8. sem images of cladocera decaying remains from sediment from lake suchar iv (poland) prepared for subfossil cladocera analysis 4 years ago. magnification: a) 500x; b) 330 x. no n c om me rci al us e o nly 182 i. zawiska et al. acknowledgments the research was founded by polish national science centre, grant ncn 2014/13/b/st10/02534 and by the eea and norway grants (grant no. 459 fss/2013/iic/w/0022). the presented work was made possible thanks to the support of institute of geography and spatial organization and institute of geological sciences, polish academy of sciences. references alonso m, 1996. [crustacea, branchiopoda].[in spanish]. museo nacional de ciencias naturales, consejo superior de investigaciones cientificas, madrid, spain. amsinck sl, jeppesen e, landkildehus f, 2005. relationships between environmental variables and zooplankton subfossils in the surface sediments of 36 shallow coastal brackish lakes with special emphasis on the role of fish. j. paleolimnol. 33:39-51. andrade-morraye m, eskinazi-sant’anna em, rocha o, 2004. a method for preparing remains of cladocera (crustacea) and chironomid (diptera: insecta) for scanning electron microscopy. braz. j. biol. 64:911-912. battarbee rw, 2001. diatoms. in: j.p. smol, b. birks john and w.m. last (eds.), tracking environmental change using lake sediments. 3. terrestrial, algal, and siliceous indicators. kluwer academic publishers, dodrecht. beyens l, meisterfeld r, 2001. protozoa. testate amoebae. in: j.p. smol, b. birks john and w.m. last (eds.), tracking environmental change using lake sediments. 3. terrestrial, algal, and siliceous indicators. kluwer academic publishers, dodrecht. chen g, dalton c, taylor d, 2010. cladocera as indicators of trophic state in irish lakes. j. paleolimnol. 44:465-481. chen wy, tao s, adams jm, jacques fmb, ferguson dk, zhou z-k, 2014. large-scale dataset from china gives new insights into leaf margin-temperature relationships. palaeogeogr. palaeoclimatol. palaeoecol. 402:73-80. duigan ca, 1992. the ecology and distribution of the littoral freshwater chydoridae (branchiopoda, anomopoda) of ireland, with taxonomic comments on some species. hydrobiologia 241:1-70. frey dg, 1986. cladocera analysis, p. 667-692. in: b.e. berglund (ed.), handbook of holocene palaeoecology and palaeohydrology. john wiley & sons, chichester. gałka m, tobolski k, zawisza e, goslar t, 2014. postglacial history of vegetation, human activity and lake-level changes at jezioro linówek in northeast poland, based on multiproxy data. veg. hist. archaeobot. 23:123-152. goldstein j, newbury de, joy dc, lyman ce, echlin p, lifshin e, sawyer l, michael jr, 2003. introduction, p. 1-20. in: j. goldstein, d.e. newbury, d.c. joy, c.e. lyman, p. echlin, e. lifshin, l. sawyer, j.r michael (eds.), scanning electron microscopy and x-ray microanalysis. 3rd ed. springer, boston. jiang w, pan h, wang f, jiang m, deng x, li j, 2015. a rapid sample processing method to observe diatoms via scanning electron microscopy. j. appl. phycol. 27:243-248. juračka pj, sacherová v, dobiášovská i, bovšková d, novosadová z, kořínek v, petrusek a, 2016. simplification of preparation techniques for scanning electron microscopy of cladocera: preparing filtering limbs and ephippia for efficient studies of ultrastructure. crustaceana 89:47-62. kemp aes, dean j, pearce rb, pike j, 2001. recognition and analysis of beddingand sediment fabric features, p. 7-22. in: w.m. last and j.p. smol (eds.), tracking environmental change using lake sediments: physical and geochemical methods. springer, dordrecht. kirillova iv, van der plicht j, gubin sv, kotov aa, 2016. taphonomic phenomenon of ancient hair from glacial beringia: perspectives for palaeoecological reconstructions. boreas 45:455-469. korhola a, 1999. distribution patterns of cladocera in subarctic fennoscandian lakes and their potential in environmental reconstruction. ecography 22:357-373. korhola a, rautio m, 2001. cladocera and other branchiopod crustaceans, p. 5-41. in: w.m. last and j.p. smol (eds.), tracking environmental change using lake sediments. 4. zoological indicators. springer, dordrecht. korhola a, tikkanen m, weckström j, 2005. quantification of holocene lake-level changes in finnish lapland using a cladocera lake depth transfer model. journal of paleolimnology 34:175-190. korosi jb, smol jp, 2012. an illustrated guide to the identification of cladoceran subfossils from lake sediments in northeastern north america: part 1 the daphniidae, leptodoridae, bosminidae, polyphemidae, holopedidae, sididae, and macrothricidae. j. paleolimnol. 48:571-586. kotov aa, 2013. morphology and phylogeny of the anomopoda (crustacea: cladocera). kmk scientific publishers, moscow: 638 pp. laforsch c, tollrian r, 2000. a new preparation technique of daphnids for scanning electron microscopy using hexamethyldisilazane. arch. hydrobiol. 149:587-596. locke a, sprules wg, 2000. effects of acidic ph and phytoplankton on survival and condition of bosmina longirostris and daphnia pulex. hydrobiologia 437:187-196. lotter a, birks hj, hofmann w, marchetto a, 1997. modern diatom, cladocera, chironomid, and chrysophyte cyst assemblages as quantitative indicators for the reconstruction of past environmental conditions in the alps. i. climate. j. paleolimnol 18: 395-420. martín-serrano a, garcía-cortés a, galán l, gallardo-millán jl, martín-alfageme s, rubio fm, ibarra pi, granda a, pérezgonzález a, garcía-lobón jl, 2009. morphotectonic setting of maar lakes in the campo de calatrava volcanic field (central spain, sw europe). sediment. geol. 222:52-63. mirosław-grabowska j, zawisza e, 2013. late glacial-early holocene environmental changes in charzykowskie lake (northern poland) based on oxygen and carbon isotopes and cladocera data. quatern. int. 328-329:156-166. nandini s, silva-briano m, garcía garcía g, sarma sss, adabache-ortiz a, galván de la rosa r, 2009. first record of the temperate species daphnia curvirostris eylmann, 1887 emend. johnson, 1952 (cladocera: daphniidae) in mexico and its demographic characteristics in relation to algal food density. limnology 10:87-94. nevalainen l, luoto tp, kultti s, sarmaja-korjonen k, 2013. no n c om me rci al us e o nly exploring the world of micro sculptures subfossil cladocera remains under the sem 183 spatio-temporal distribution of sedimentary cladocera (crustacea: branchiopoda) in relation to climate. j. biogeogr. 40:1548-1559. nevalainen l, rautio m, 2014. spectral absorbance of benthic cladoceran carapaces as a new method for inferring past uv exposure of aquatic biota. quaternary sci. rev. 84:109-115. nevalainen l, sarmaja-korjonen k, luoto tp, 2011. sedimentary cladocera as indicators of past water-level changes in shallow northern lakes. quaternary res. 75:430-437. rumes b, eggermont h, verschuren d, 2011. distribution and faunal richness of cladocera in western uganda crater lakes. hydrobiologia 676:39-56. sinev ay, elmoor-loureiro lm, 2010. three new species of chydorid cladocerans of subfamily aloninae (branchipoda: anomopoda: chydoridae) from brazil. zootaxa 2390:1-25. sinev ay, van damme k, kotov a, 2005. redescription of tropical-temperate cladocerans alona diaphana king, 1853 and alona davidi richard, 1895 and their translocation to leberis smirnov, 1989 (branchiopoda: anomopoda: chydoridae). arthropoda selecta 14:183-205. sousa fd, elmoor-loureiro lm, debastiani-junior jr, mugnai r, senna a, 2015. new records of anthalona acuta van damme, sinev & dumont 2011 and anthalona brandorffi (sinev & hollwedel, 2002) in brazil, with description of a new species of the simplex-branch (crustacea: cladocera: chydoridae). zootaxa 4044:224-240. szeroczyńska k, sarmaja-korjonen k, 2007. atlas of subfossil cladocera from central and northern europe. friends of the lower vistula society: 84 pp. van damme k, dumont hj, 2007. limb morphology of the carnivorous anomopods anchistropus emarginatus sars, 1862 and pseudochydorus globosus (baird, 1843) (crustacea: branchiopoda: anomopoda). ann. limnol-int. j. lim. 43:271-284. wetzel a, 2013. formation of methane-related authigenic carbonates within the bioturbated zone an example from the upwelling area off vietnam. palaeogeogr. palaeoclimatol. palaeoecol. 386:23-33. zawiska i, słowiński m, correa-metrio a, obremska m, luoto t, nevalainen l, woszczyk m, milecka k, 2015. the response of a shallow lake and its catchment to late glacial climate changes a case study from eastern poland. catena 126:1-10. zawiska i, zawisza e, woszczyk m, szeroczyńska k, spychalski w, correa-metrio a,2013. cladocera and geochemical evidence from sediment cores show trophic changes in polish dystrophic lakes. hydrobiologia 715:181-193. no n c om me rci al us e o nly layout 1 introduction hydrological alterations can deeply modify the biotic composition of aquatic ecosystems (richter et al., 1996). this is particularly true in those water bodies used to fulfil agriculture needs and located in semi-arid regions (naselliflores, 2003). the extent of these operational procedures can be strong enough to cause a significant reduction in the stored volumes (jeppesen et al., 2015) and, in some cases, the complete drying out of shallow lakes and wetlands. as a consequence of volume loss, water bodies can experience deep changes in their chemical and physical characteristics; local extinction of taxa and changes in the biological structure of aquatic communities due to species replacements may therefore occur. calanipeda aquaedulcis is the only pseudodiaptomid copepod occurring in the inland waters of the west palearctic biogeographical region, and the only species of the monotypic genus calanipeda kritschagin, 1873 (dussart and defaye 2002). this widespread copepod is known to occur from the circum-mediterranean area to the black and caspian seas regions, where it occurs mostly in brackish, coastal, large water bodies (błędzki and rybak 2016), although it is also known to occur in freshwater lakes (marrone et al., 2019). in italy, c. aquaedulcis is reported for several localities throughout the country (see cannicci 1939; stella 1984; ruffo and stoch 2005; belmonte, 2018); conversely, a single population was to date reported for sicily: the one occurring in the shallow lake biviere di gela (stella 1979, 1984; margaritora et al., 1982; calvo et al., 1993; ruffo and stoch 2005). at the end of the xx century, the lake, located at 1.3 km from the southern sicilian coast and 8 m asl, underwent a period of severe water reduction which caused the intrusion of seawater with a consequent increase in its conductivity values (jeppesen et al., 2015). these events caused the disappearing of aquatic macrophytes and the recurrent massive development of the toxin-producing haptophyte prymnesium parvum carter (naselli-flores and barone, 2019), known to be detrimental for other planktonic organisms due to its toxins (fistarol et al., 2003). these blooms deeply changed the biological structure of the lake, and barone et al. (2010) recorded the extinction of the crustacean zooplankton inhabiting the lake, which likely occurred during the first blooms of p. parvum. among those species, the only known sicilian population of calanipeda aquaedulcis had vanished (calvo et al. 1993 and literature therein). in this paper, we present some selected results from an extensive sampling campaign conducted on both temporary and permanent water bodies throughout sicily. this allowed checking whether additional populations of the species were present on the island or if calanipeda aquaedulcis should be considered locally extinct. methods samples were collected from 2003 to 2019 in more than 700 water bodies spread throughout mainland sicily (f. marrone, unpublished data). permanent and temporary natural ponds and swamps, agriculture artificial ponds, and larger dam-reservoirs were sampled. the latter ecosystem type had already been sampled in 1987/88 by calvo et al. (1993). some of the samples collected by these last authors are still available at the university of palermo and were therefore re-analyzed to check eventual differences in the species composition with the more recently collected samples. in each visited water body, water temperature and short note on the occurrence and distribution of calanipeda aquaedulcis kritschagin, 1873 (copepoda, calanoida, pseudodiaptomidae) in sicily, italy, with some notes on coexistence and species replacement in calanoid copepods luca vecchioni, federico marrone, luigi naselli-flores department of biological, chemical and pharmaceutical sciences and technologies, university of palermo, via archirafi 18, 90123 palermo, italy abstract the only population of the pseudodiaptomid copepod calanipeda aquaedulcis kritschagin, 1873 to date reported to occur in sicily disappeared at the beginning of the xxi century due to deep environmental changes which affected the single site (lake biviere di gela) known for this species on the island. in that site c. aquaedulcis is now replaced by copidodiaptomus numidicus (gurney, 1909), a diaptomid copepod whose distribution has been greatly increasing since the second half of the last century. in the present note, the occurrence of c. aquaedulcis in 12 novel water bodies spread throughout sicily is reported, and some environmental data on the sites where the species was collected are provided. moreover, in a few sites c. aquaedulcis was found to co-occur with copidodiaptomus numidicus. the recorded co-occurrence of these two species and the replacement of c. aquaedulcis with c. numidicus in lake biviere di gela are briefly discussed. no nco mm er cia l u se on ly calanipeda aquaedulcis in sicily 19 electric conductivity at 20°c were measured with a hanna instruments hi9835 multiprobe. due to the small dimensions of the majority of sampled sites, plankton was collected from the shoreline with a 125 micron meshsized plankton net; in the largest water bodies, vertical plankton hauls were also performed from a boat in the middle of the lake. collected samples were fixed in situ in 80% ethanol; calanoid copepods were sorted in the laboratory and identified to species level according to kiefer (1978) and stella (1984). all the samples are stored at the university of palermo, italy, in the crustacean collection of fm. results in twelve out of the 728 sampled water bodies, populations of calanipeda aquaedulcis were found (tab. 1, fig. 1). the species was also present in the zooplankton samples collected by calvo et al. (1993) on 4th september and 2nd december 1987 in lake disueri (cl024), although the presence of this species, likely by mistake, was not included in the checklist of the zooplankton of the lake provided by these authors. conversely, in the sample collected in the same lake on 10th april 2019 by the authors of the present note, no calanoid copepods were present. in lake cimia (cl027), where the presence of the diaptomid copidodiaptomus numidicus had been reported by calvo et al. (1993), the co-existence of c. aquaedulcis and c. numidicus was observed on 10th april 2019 (tab. 1). although when present calanipeda aquaedulcis was, in the majority of cases, the only calanoid copepod recorded, it synchronically co-occurred with copidodiaptomus numidicus in four sites (tab. 1). these two species were generally co-occurring at conductivity values below 5 ms cm–1 (cl024, cl027, cl028 and en004), even though in a single site (cl028, a farm pond close to lake cimia, filled with waters from that lake) both species were present at conductivity values above 14 ms cm–1. the sites with c. aquaedulcis as the only calanoid species were a few brackish coastal waterbodies (tp107, tp114, tp132, tp133, tp134 and tp165) located on the western coast of sicily and characterized by conductivity values above 11 ms cm–1 due to seawater intrusions. in addition, it was the only calanoid species in a brackish coastal swamp (sr022) by the ionian sea (conductivity above 3 ms cm–1), and in the lake biviere di lentini (sr011), a dam reservoir filled in the mid of the ‘90s of the last century and characterized by conductivity values ranging between 1.3 and 1.7 ms cm–1. discussion human disturbance and the building of artificial water bodies are known to have significant effects on ta b. 1 .l is t o f t he o cc ur re nc e si te s of c al an ip ed a aq ua ed ul ci s in s ic ily , i ta ly . d ec im al g eo gr ap hi ca l c oo rd in at es a re re po rt ed a cc or di ng to w g s8 4 m ap d at um . c od e s it e na m e l at it ud e n l on gi tu de e a lt it ud e h ab it at ty pe h yd ro pe ri od sa m pl in g e c t em p. o cc ur ri ng ( m a sl ) d at e (m s cm –1 ) (° c ) c al an oi da c l 00 6 b iv ie re d i g el a 37 ,0 19 72 1 1 4, 34 44 34 8 s ha llo w la ke p er m . b ef or e 20 05 2 .4 -2 .9 1 2. 629 .0 c . a qu ae du lc is * c l 02 4 l ag o d is ue ri 3 7, 19 25 06 14 ,2 86 71 3 1 51 d am re se rv oi r p er m . 04 /0 9/ 19 87 1 .6 2 9. 0 c . n um id ic us ; c . a qu ae du lc is § c l 02 7 l ag o c im ia 37 ,1 89 83 0 1 4, 35 42 17 1 37 d am re se rv oi r p er m . 10 /0 4/ 20 19 1 .5 1 6. 7 c . n um id ic us ; c . a qu ae du lc is ‡ c l 02 8 st ag no c im ia 37 ,1 86 46 1 1 4, 35 65 10 1 26 f ar m p on d n .a . 10 /0 4/ 20 19 1 4. 5 2 1. 6 c . n um id ic us ; c . a qu ae du lc is e n 00 4 m as se ri a d io da to p 2 37 ,5 55 39 2 1 4, 72 19 81 1 41 f ar m p on d p er m . 15 /0 9/ 20 07 4 .5 2 9. 6 c . n um id ic us ; c . a qu ae du lc is t p1 07 c ap o fe to p 1 37 ,6 61 28 4 1 2, 52 70 90 3 s w am p p er m . 25 /0 2/ 20 07 3 8. 6 1 9. 3 c . a qu ae du lc is t p1 14 m ar gi o sp an ò p2 37 ,6 78 51 6 1 2, 50 02 82 3 s w am p t em p. 20 /0 3/ 20 14 3 8. 0 2 2. 0 c . a qu ae du lc is t p1 32 c ap o fe to p 5 37 ,6 61 76 3 1 2, 53 31 61 4 c ha nn el pe rm 1 2/ 01 /2 00 9 3 2. 5 1 3. 0 c . a qu ae du lc is t p1 33 c ap o fe to p 6 37 ,6 58 93 0 1 2, 53 43 09 4 p on d p er m . 12 /0 1/ 20 09 3 0. 9 1 3. 3 c . a qu ae du lc is t p1 34 c ap o fe to p 7 37 ,6 60 28 4 1 2, 53 82 66 4 m ar sh te m p. 12 /0 1/ 20 09 2 8. 4 1 2. 9 c . a qu ae du lc is t p1 65 m ar gi o m ilo p 2 3 7, 72 42 39 12 ,4 74 09 3 1 s w am p p er m 2 3/ 12 /2 01 1 1 1. 0 1 3. 0 c . a qu ae du lc is sr 01 1 b iv ie re d i l en tin i 3 7, 31 49 53 14 ,9 34 89 3 2 9 d am re se rv oi r p er m . 29 /0 1/ 20 03 ≈ 1. 5* * n. d. c . a qu ae du lc is sr 02 2 pa nt an o g el sa ri 3 7, 33 83 13 15 ,0 85 73 0 0 s w am p t em p. 01 /0 4/ 20 07 3 .1 2 1. 0 c . a qu ae du lc is * i n th e si te c l 00 6, c . a qu ae du lc is w as r ep la ce d by c op id od ia pt om us , n um id ic us fr om 2 00 5 on w ar ds ( se e te xt ); § i n th is s ite th e sp ec ie s w as f ou nd a ls o on 2 nd d ec em be r 19 87 b ut n ot o n 10 th a pr il 20 19 ; ‡ in th is s ite , c al vo e t a l. (1 99 3) c ol le ct ed o nl y c . n um id ic us in 1 98 719 88 ; * * d at a fr om r eg io ne s ic ili an a, 2 00 7b ; p er m ., pe rm an en t h yd ro pe ri od ; t em p. , t em po ra ry h yd ro pe ri od ; n .a .: no t a va ila bl e; n .d .: no t d et ec te d. no nco mm er cia l u se on ly l. vecchioni et al.20 species distribution, favoring the spreading and establishment of the more euryecious species at the expense of the more stenoecious ones (miracle, 1982; parkes and duggan, 2012; riley et al., 2018). such a process leads to an alteration of the original distribution patterns, to the homogenization of faunas at a regional and global scale (rahel, 2002), and to the regression or extinction of the native communities, which are often of great biogeographic and conservation value (e.g. the “hemidiaptomus ponds”, sensu sahuquillo and miracle, 2013). moreover, water movements among different water bodies due to management procedures commonly occur in semi-arid areas (marrone and naselli-flores, 2019, in press). these are often drastic and unpredictable, and do not keep in any account the alterations they can cause on the structure of biological communities (naselli-flores, 1999). the stress caused by humandriven operational procedures in man-made lakes can eventually lead to local extinction of species and contribute to weaken those biological filters which act as a barrier against the colonization of new species (incagnone et al., 2015). local species extinction can open the way to more euryecious taxa, and eventually to invasive species. in accordance to what reported by barone et al. (2010) and naselli-flores (unpublished data), calanipeda aquaedulcis was not found in the samples collected in lake biviere di gela from 2003 onward (until april 2019); fig. 1. occurrence localities of calanipeda aquaedulcis in sicily. black circles, occurrence of c. aquaedulcis; black-and-red circles, co-existence of c. aquaedulcis and copidodiaptomus numidicus; wwhite circle, previous presence of c. aquaedulcis, now replaced by c. numidicus; yellow circle, co-occurrence of c. aquaedulcis and c. numidicus in 1987, no calanoids species in 2019. no nco mm er cia l u se on ly calanipeda aquaedulcis in sicily 21 in this lake, c. aquaedulcis is now replaced by the euryecious diaptomid copepod copidodiaptomus numidicus, which was first detected in 2005. the disappearing of c. aquaedulcis from lake biviere di gela likely occurred at the end of the last century when the lake depth reached values ranging between 1.2 and 1.5 m due to the intensification of water demand for irrigation purposes (jeppesen et al., 2015). this decrease in water levels caused drastic changes in the chemical, physical and biological features of the lake, which allowed the establishment of the toxic haptophyte prymnesium parvum, and made this environment unsuitable for the pseudodiaptomid (for more details see barone et al., 2010; jeppesen et al., 2015). the re-analysis of some of the samples collected by calvo et al. (1993) revealed that calanipeda aquaedulcis was present with low numbers in lake disueri in the ‘80s, whereas the species was not observed in the samples collected in the same lake in spring 2019. lake disueri is a dam reservoir located 20 km north west of lake biviere di gela. the lake, since the mid 1990s, has been connected by a pipeline to lake cimia to fulfill irrigation needs. actually, although both the lakes were built in the catchment of gela river, they rest on two parallel river stretches. these merge downstream both the dams, before flowing into the sicilian channel, about 8 km from the lake biviere di gela. all these lakes have conductivity values higher than 1.5 ms cm–1 and show chloride-sulphate-alkalineearth waters, even though the chemical characteristics of lake biviere di gela are due to the intrusions of seawater (barone et al., 2010) whereas those of lake disueri and lake cimia are due to the outcrops of evaporites in their catchment (calvo et al., 1993). at the end of the 1980s, the capacity of lake disueri had already been drastically reduced by siltation (calvo et al., 1993), and in 2006 the lake was completely dried out in order to carry out dam maintenance (regione siciliana, 2007a). the absence of calanoid copepods in the samples collected on april 10th, 2019 in lake disueri is probably to be ascribed to the several episodes of total desiccation the lake underwent in the past years, which likely caused a re-arrangement of its biological structure. at the same time, the pipeline connecting lake disueri to lake cimia may have favored the dispersal of c. aquaedulcis in the latter water body and, from there, in the agriculture pond fed with lake cimia’s waters. in spite of the extinction of the population inhabiting the lake biviere di gela, calanipeda aquaedulcis proved to be a rather infrequent but widespread species in sicily. in good accordance with the available literature, most of its occurrence localities are permanent, coastal, brackish water bodies. nevertheless, the species was also collected in temporary ponds and swamps (e.g. sr022, tp114, and tp134), and even in man-made reservoirs located in the sicilian hinterland (e.g. sr011, en004, cl024, cl027 and cl028). these latter sites, which are located up to 33 km far from the seashores and up to 151 m above sea level, are characterized by significantly mineralized but atalassohaline waters, further stressing the euryecious character of the species. in fact, as already highlighted by cannicci (1939) and borutzky et al. (1991), c. aquaedulcis seems to be an euryhaline species thriving in poorly mineralized to isohaline waters, and its occurrence in hypersaline water bodies is to be considered only sporadic. even though gonçalves et al. (2012) reported the coexistence of c. numidicus and c. aquaedulcis in the river mondego (portugal), where they could be possibly part of the biological drift, the synchronic co-occurrence of adult stages of c. aquaedulcis and c. numidicus in lentic water bodies is to our knowledge unprecedented. it would be therefore worth investigating the phenology of these assemblages to check whether the two species can co-exist in the long term or are rather experiencing a process of species replacement as already observed in lake biviere di gela. conclusions the sampling campaign carried out in sicilian inland waters so far has shown that, although sound synoptic data are available for some taxonomical groups (e.g. diaptomids, see marrone et al., 2017), the biodiversity assessment of mediterranean aquatic ecosystems is still far from being complete. several circum-mediterranean regions have not been adequately investigated yet, as well as many large islands in the mediterranean basin (ruffo and stoch, 2005; marrone, 2006). this makes difficult to assess the actual biodiversity of mediterranean inland waters, the appreciation of its value, and the zoogeographical relationships existing among the different faunas. at the same time, the disappearance of rare aquatic species and the spreading of alien, invasive ones can be overlooked due the paucity or lack of information (marrone and naselli-flores, 2015), further increasing the threats to which these ecosystems are subjected. finally, in the light of the wide distribution of c. aquaedulcis and of the frequent cryptic speciation in all the calanoid copepod families (marrone et al., 2013; cornils and held 2014; sabia et al., 2017; kasapidis et al., 2018, and references therein), the carrying out of a phylogeographic study of the species throughout its known distribution range is desirable. no nco mm er cia l u se on ly corresponding author: federico.marrone@unipa.it keywords: crustacea; zooplankton local extinction; species replacement; mediterranean basin. conflict of interest: the authors declare no competing interests. ethical statement: this article does not contain any studies with human participants or animals performed by any of the authors. received: 15 march 2019. accepted: 21 may 2019. this work is licensed under a creative commons attribution noncommercial 4.0 license (cc by-nc 4.0). ©copyright: the author(s), 2019 licensee pagepress, italy advances in oceanography and limnology, 2019; 10:8177 doi: 10.4081/aiol.2019.8177 references barone r, castelli g, naselli-flores l, 2010. red sky at night cyanobacteria delight: the role of climate in structuring phytoplankton assemblage in a shallow, mediterranean lake (biviere di gela, southeastern sicily). hydrobiologia 639:43-53. belmonte g, 2018. calanoida (crustacea: copepoda) of the italian fauna: a review. eur. zool. j. 85:274-290. błędzki la, rybak ji, 2016. freshwater crustacean zooplankton of europe. springer, switzerland: 923 pp. borutzky eb, stepanova la, kos ms, 1991. opredelitel’ calanoida presnykh vod sssr (key to the fresh-water calanoids of the urss). nauka, st. petersburg: 504 pp. calvo s, barone r, naselli-flores l, fradà orestano c, dongarrà g, lugaro a, genchi, g, 1993. limnological studies on lakes and reservoirs of sicily. nat. sicil. 27(suppl.):1-292. cannicci g, 1939. sulla distribuzione della popella guernei richard (calanipeda aquae dulcis kritz.) negli stagni salmastri. int. rev. hydrobiol. 38:489-503. dussart b, defaye d, 2002. world directory of crustacea copepoda of inland waters, i. calaniformes. backhuys publisher, leiden: 276 pp. fistarol g, legrand c, granéli e, 2003. allelopathic effects of prymnesium parvum on a natural plankton community. mar. ecol. prog. ser. 255:115-125. gonçalves amm, azeiteiro um, pardal ma, de troch m, 2012. fatty acid profiling reveals seasonal and spatial shifts in zooplankton diet in a temperate estuary. estuar. coast. shelf. sci. 109:70-80. jeppesen e, brucet s, naselli-flores l, papastergiadou e, stefanidis k, nõges t, nõges p, attayde jl, zohary t, coppens j, bucak t, fernandes menezes f, sousa freitas fr, kernan m, søndergaard m, beklioğlu m, 2015. ecological impacts of global warming and water abstraction on lakes and reservoirs due to changes in water level and related changes in salinity. hydrobiologia 750:201-227. kasapidis p, siokou i, khelifi-touhami m, mazzocchi mg, matthaiaki m, christou e, fernandez de puelles ml, gubanova a, di capua i, batziakas s, frangoulis c, 2018. revising the taxonomic status and distribution of the paracalanus parvus species complex (copepoda, calanoida) in the mediterranean and black seas through an integrated analysis of morphology and molecular taxonomy. j. plankton res. 40:595-605. kiefer f, 1978. [das zooplankton der binnengewässer. freilebende copepoda. die binnengewässer, band 26 teil 2].[book in german]. e. schweizerbart’sche verlagbuchhandlung, stuttgart: 343 pp. margaritora fg, mastrantuono l, crosetti d, lombardi f, 1982. contributo alla conoscenza della fauna ad entomostraci delle acque interne della sicilia. animalia 9:87-102. marrone f, 2006. the microcrustacean fauna of sicily and the central mediterrranean sea area – current knowledge and gaps to be filled. pol. j. ecol. 54:681-685. marrone f, alfonso g, naselli-flores l, stoch, f, 2017. diversity patterns and biogeography of diaptomidae (copepoda, calanoida) in the western palearctic. hydrobiologia 800:45-60. marrone f, alfonso g, stoch f, pieri v, alonso m, dretakis m, naselli-flores l, 2019. an account on the nonmalacostracan crustacean fauna from the inland waters of crete, greece, with the synonymization of arctodiaptomus piliger brehm, 1955 with arctodiaptomus alpinus (imhof, 1885). limnetica 38:167-187. marrone f, lo brutto s, hundsdoerfer ak, arculeo m, 2013. overlooked cryptic endemism in copepods: systematics and natural history of the calanoid subgenus occidodiaptomus borutzky 1991 (copepoda, calanoida, diaptomidae). mol. phylogenet. evol. 66:190-202. marrone f, naselli-flores l, 2015. a review on the animal xenodiversity in sicilian inland waters (italy). adv. oceanogr. limnol. 6:2-12. marrone f, naselli-flores l (2019). different invasibility of permanent and temporary water bodies in a semi-arid mediterranean island. inland waters (in press). miracle mr, 1982. biogeography of freshwater zooplanktonic communities of spain. j. biogeog. 9:455-467. naselli-flores l, 1999. limnological aspects of sicilian reservoirs: a comparative, ecosystemic approach, p. 283-311. in: j.g. tundisi and m. straškraba (eds.), theoretical reservoir ecology and its applications. backhuys publishers, leiden. naselli-flores l, 2003. man-made lakes in mediterranean semiarid climate: the strange case of dr deep lake and mr shallow lake. hydrobiologia 506-509: 13-21. naselli-flores l, barone r, 2019. mixotrophic phytoplankton dynamics in a shallow mediterranean water body: how to make a virtue out of necessity. hydrobiologia 831:33-41. parkes sm, duggan ic, 2012. are zooplankton invasions in constructed waters facilitated by simple communities? divers. distrib. 18:1199-1210. rahel f, 2002. homogenization of freshwater faunas. annu. rev. ecol. syst. 33:291-315. regione siciliana, 2007a. [piano di tutela delle acque della sicilia. bacino idrografico gela (r19077)].[in italian]. sogesid, palermo. accessed on: april 21, 2019. available from: http://www.osservatorioacque.it/documenti/pta/b26.pdf l. vecchioni et al.22 no nco mm er cia l u se on ly calanipeda aquaedulcis in sicily 23 regione siciliana, 2007b. piano di tutela delle acque della sicilia. bacino idrografico lentini e bacini minori tra lentini e simeto (r19093].[in italian]. sogesid, palermo. accessed on: april 21, 2019. available from: http://www.osservatorioacque.it/documenti/pta/b36.pdf richter bd, baumgartner jv, powell j, braun dp, 1996. a method for assessing hydrological alteration within ecosystems. conserv. biol. 10:1163-1174. riley wd, potter ece, biggs j, collins al, jarvie hp, jones ji, kelly-quinn m, ormerod sj, sear da, wilby rl, broadmeadow s, brown cd, chanin p, copp gh, cowx ig, grogan a, hornby dd, huggett d, kelly mg, naura m, newman jr, siriwardenat gm. 2018. small water bodies in great britain and ireland: ecosystem function, human-generated degradation, and options for restorative action. sci. total. environ. 645:1598-1616. ruffo s, stoch f, 2005. [checklist e distribuzione della fauna italiana].[article in italian]. mem. mus. civico storia nat. verona, 2 ser., sez. sci. vita 16:1-307. sabia l, di capua i, percopo i, uttieri m, amato a, 2017. its2 in calanoid copepods: reconstructing phylogenetic relationships and identifying a newly introduced species in the mediterranean. eur. zool. j. 84:104-115. sahuquillo m, miracle mr, 2013. the role of historic and climatic factors in the distribution of crustacean communities in iberian mediterranean ponds. freshwater biol. 58:1251-1266. stella e, 1979. [nuovi reperti di calanidi (crustacea copepoda) in italia].[article in italian]. boll. mus. civico storia nat. verona. 6:333-346. stella e, 1984. [copepoda: calanoida. xxi fauna d’italia]. [book in italian]. ed. calderini, bologna: 101 pp. no nco mm er cia l u se on ly layout 1 introduction water characteristics and quality are defined by specific physical, chemical and biological properties, and how these properties impact the survival, reproduction, growth and management of aquatic life (aduwo and adeniyi, 2019). indeed, the sustainability and development potential of any country may not be possible without water (salla and ghosh, 2014). lakes may be monitored for recreational, domestic and/or used as a component of hydro-power generating systems. in developing countries, lakes are primarily used by the local inhabitants for transportation, fishing, washing, cooking and irrigation practices (okoro et al., 2014). however, most records these days indicates that water quality is increasingly deteriorating, and this is cause of a global concern (mahananda et al., 2005). a previous study (salla and ghosh, 2014) emphasized that about 75% of surface water may be contaminated by different kinds of pollutants. pollutants may include heavy metals (awoyemi et al., 2014) or nutrients (e.g. phosphate and nitrate) from industrial discharges (mahananda et al., 2005) or agricultural activities. higher concentrations of nutrients, e.g. phosphorus and nitrogen, may cause hypoxia and algal bloom (anitha, 2002; garg et al., 2009), which may cause low light penetration, obstruction of oxygen levels and loss of aquatic life and biodiversity (e.g. fish mortality) (parashar et al., 2006). for these reasons, the constant monitoring of water quality in lakes is essential and can be conducted by quantifying the level of physical and chemical parameters. the physical properties of a lake include water temperature, colour, odour and taste, solids and turbidity (aduwo and adeniyi, 2019), while the chemical assessment involves the measurements of cations, anions, nutrient compounds, toxic and non-toxic compounds, and oxygen demands by inorganic and organic substances (ademoroti, 1996). a typical lake and a major source of water for the local population of south-eastern and niger-delta region in nigeria is lake oguta. the lake serves as a source of income to the community (fishing and also dredging the lake for sand, which is used in the construction industries) (nfor and akaegbobi, 2012). however, the lake has been exposed to threats due to excessive oil exploration (isinkaye and emelue, 2015) and sewage disposal (nfor and akaegbobi, 2012). currently, records on the physical and chemical status of the lake are needed, but very few are existing or have been explored (nfor and akaegbobi, 2012). the investigation of water quality levels (physical and chemical properties) of lake oguta by a comparison with a set of standards (dirican, 2015) may provide information for government management policies (patil et al., 2012). the objective of this paper is to evaluate water quality and potential human exploitation (including drinking purposes) of lake oguta using a suitable subset of variables measured at weekly intervals. the parameters article assessment of the physical and chemical properties of lake oguta (nigeria) in relation to the water quality standard established by the nigerian federal ministry of water resources felix atawal andong1,2*, ngozi evelyn ezenwaji1, temitope dadewura melefa1, funmilayo faith hinmikaiye1, obiechina vitus nnadi1, olasoji oluwafemi1 1departmentof zoology and environmental biology, university of nigeria, nsukka, enugu state; 2 a.p. leventis ornithological research institute, department of zoology, university of jos, plateau state, nigeria abstract constant assessment of physical and chemical parameters in freshwater ecosystems is largely recommended. this is even more important when water resources, e.g. lakes in most countries, serve as a source of water for domestic and commercial purposes, and /or when freshwater ecosystems represent a refuge for most aquatic organisms. in this paper, we investigated the physical and chemical properties of water resources at three sampling stations of lake oguta, comparing the weekly values (june-july 2018) with the water quality standard established by the nigerian federal ministry of water resources (fmwr). the parameters analyzed included water temperature, ph, dissolved oxygen (do), chemical and biological oxygen demand (cod, bod), potassium, magnesium, sodium, calcium, phosphate, nitrate, chloride and sulphate. most of the cations (calcium, magnesium and sodium), anions (phosphate, nitrate, chloride and sulphate), as well as water temperature, bod and do were below the quality standard limits. the basic chemistry and temporal variations may have been caused mostly by natural factors such as geology, topography, meteorology, hydrology, water levels and biological activity. being in line with the recommended standard levels, the nutrient concentrations, ph and hardness in the current study may indicate favourable conditions for the life of aquatic organisms and contemporary co-existence with the human exploitation for drinking purposes. nevertheless, to assure a safely and conscious exploitation of this water resource, we recommend continuity in the monitoring studies. to assure an accurate evaluation of the physical and chemical parameters, future studies should include a larger sample size and extended study periods (including other seasons). no nco mm er cia l u se on ly water quality of lake oguta, nigeria 75 that have been measured, i.e. water temperature, ph, dissolved oxygen, chemical and biological oxygen demand, potassium, magnesium, sodium, calcium, phosphate, nitrate, chloride and sulphate were compared with the standards assessed by the nigerian federal ministry of water resources fmwr (nis, 2015). methods study area imo state is located in south-eastern nigeria; currently, the state is famous for its largest natural lake or fresh water lake, lake oguta (nfor and akaegbobi, 2012) (fig. 1). the lake is located in a low-lying platform, at 50 m above sea level, between latitudes 5°4’’and 5°44’ n, and longitudes 6°45’and 6°50’ e. four rivers (njaba, awbuna, utu and orashi) are connected to the lake (ahiarakwem and onyekuru, 2011). however, all year round, rivers njaba and awbuna discharges into the lake, while utu stream flows in during the rainy season (ahiarakwem et al., 2012). the river orashi flows past the lake in its south-western portion. it is recorded that, the total annual inflow from the rivers and streams is about 25,801 m3 (ahiarakwem, 2006), while the annual return and overland flow into the lake is estimated to be about 69,000 and 138,000 m3 (okoro et al., 2014). also, the annual recharge of the lake from precipitation is about 693,000 m3, while the annual groundwater inflow into the lake has been estimated (2,750,400 m3) (ahiarakwem et al., 2012). indeed, the total annual water inflow, heavily outweighs the total annual outflow, thus, the lake is adequately recharged all the year round (ahiarakwem, 2006). the surface area of the lake ranges between 1.8 km2 and 2.5 km2. the shoreline length is around 10 km; maximum and mean depths are 8.0 m and 5.5 m, fig. 1. map of lake oguta. sampling stations are indicated with red triangles. no nco mm er cia l u se on ly f.a. andong et al.76 respectively (nfor and akaegbobi, 2012). may be, owing to the low cation concentration and alkalinity of the water, conductivity may range between 8.6-30 µs cm–1 (odigi and nwadiaro, 1988; okorondu and anyadoh-nwadike, 2015); also, the surface water temperature (24-31°c), secchi-disc-transparency (0.61-4.50 m) (odigi and nwadiaro,1988; nwadiaro, 2018) may vary seasonally with rainfall. the lake experiences a weak, unstable thermal stratification, which may develop towards the middays of the warmer months (i.e. april-august) (nwadiaro, 2018). selection of sampling site, data collection and processing after a pre-sampling survey aimed at assessing the general characteristics of the lake and activities in the catchment, we identified three peculiar zones, i.e. station 1, or lake bank (human activities active, e.g. washing); station 2, lake flows into each other; and station 3, area with active farming. we traversed the lake using a boat and collected water samples every week, from june to july, 2018. the distance between sampling stations was 2.5 km; and before sample collection, we ensured that the containers were thoroughly rinsed with the lake water. during our sample collection we were guided by a standard procedure (apha, 1998; awoyemi et al., 2014). water samples were collected by gently lowering the container into the lake (ozoko, 2015). in the field, each water sample was first analysed to quantify dissolved oxygen (do) using the winkler’s method (aduwo and adeniyi, 2019). we measured also, the surrounding temperature (28.0-33.0°c) and water temperature (25.527.8°c) for each station using a mercury-in-bulb thermometer (graduation 0-360°c). the probe was lowered into the water for at least five minutes, and readings were recorded immediately while the thermometer was still in the water, necessarily to avoid interference with ambient temperature. hydrogen activity (ph) was measured using a portable hanna field ph meter (model phs 25), inserted into the water and allowed to attain a steady value. further, water samples for chemical analyses were preserved using 10 ml 6n hno3 (awoyemi et al., 2014), and transported to the national center for energy research and development laboratory, university of nigeria, nsukka, enugu state (6º 51’21” n 7º23’45” e). in the laboratory, biochemical oxygen demand (bod) was quantified using iodiometric titration (aduwo and adeniyi, 2019). while chemical oxygen demand (cod) was quantified in the laboratory using the chromic acid wet digestion titrimetric method (awoyemi et al., 2014). an ultraviolet atomic absorption spectrophotometer (ozoko, 2015), was used for quantification of cations, i.e. sodium (na+), potassium (k+), magnesium (mg2+) and calcium (ca2+). while the major anions, i.e. phosphate (po43-), nitrate (no3-), sulphate (so42-), were assessed through ion chromatography (dionex) (apha, 1998; dirican, 2015). statistical analysis data were analysed using spss (statistical package for social science) version 20.0; the physical and chemical parameters were first checked for normality and homogeneity of variance. analysis of variance (anova) was used to test the level of significance (set at p<0.05); duncan’s new multiple range test (dnmrt) was used to separate monthly station means. results the results of the physical and chemical parameters quantified from the analysis of lake oguta, compared to water quality standard established by the nigerian federal ministry of water resources (fmwr), are presented in tab. 1. water temperature was significantly different between months in two stations, i.e. stations 2 and 3 (p=0.01 and p=0.03, respectively). the range of values in the three stations (25.6-27.8°c), was below the federal standard limit (<35ºc). values of ph were significantly different between months in station 3 (p=0.03); the ph range in the current study (6.5-7.1) was within the standard limit (6.5-8.5). also, do range was below (3.45.7 mg l– 1) the standard limit (7.5 mg l–1), and was not statistically significant between months in the three stations (p>0.05). cod values ranged between 24.4 and 26.7 mg l–1; differences between months were statistically significant only in station 3 (p=0.01). bod was significant between months in stations 1 and 3 (p=0.01 for both); further, the range of bod values recorded in the three stations (3.9-5.9 mg l–1) was below the standard limit (8.3 mg l–1). potassium (2.6-4.3 mg l–1) showed significant differences between months in stations 1 and 2 (p=0.01). sodium (1.32.5 mg l–1) showed significant differences in station 3 (p=0.03); below the standard limit (200 mg l–1). among algal macronutrients, i.e. phosphate and nitrate; phosphate concentrations (0.9-1.3 mg l–1) showed no differences between months, and having lower values compared to the standard limit (<13.5). while nitrates ranged between 2.4 and 5.8 mg l–1 below the standard limit (50 mg l–1) with significant differences between months in station 3 (p=0.01). other parameters ranged between 10.5 and 14.0 mg l–1 (chloride) and 3.5 and 5.6 mg l–1 (sulphate). at the stations 1 and 2, differences in chloride concentrations were statistically significant (p<0.05), while sulphate was significant only in station 1 (p=0.04); these two parameters were below the standard limit. the water parameters measured, e.g. magnesium and calcium especially in its complex form, i.e. mgcaco3 (150 mg l–1), is used as an no nco mm er cia l u se on ly water quality of lake oguta, nigeria 77 indicator of water hardness in nigeria (nis, 2007). magnesium concentrations were significantly different in station 3 (p=0.01) and ranged between 3.1 and 4.6 mg l–1, below the standard limit (20 mg l–1), while calcium values were not statistically different between months in any of the three station, and ranged between 1.9 and 2.9 mg l–1, below the standard limit (150 mg l–1). discussion in the current study, the range of surface water temperature in the three stations was not beyond the federal standard (<35°c) and also within the values reported in other studies, e.g. 26.5-33°c (oluyemi et al., 2010); 26.4-31°c (rim-rukeh, 2013); 25-28°c (who, 2011); and 24.2-26.2°c (nwoko et al., 2015). usually, air temperature can seriously influence water temperature, hereby causing differences in the surface and / or midbottom level temperature of a lake. such changes in water temperature can also influence other water quality indices (dirican, 2015). in the current study, the weather condition was cooler due to several rain showers experienced at the sampling station. high temperature values increase the metabolic rate of aquatic organisms, with important effects on o2 production and consumption. in case of excessive primary production, this can cause a tab. 1. monthly variations in the physical and chemical parameters in the three sampled stations, and comparison with the limits set by the standards of the nigerian federal ministry of water resources. variable duration station 1 station 2 station 3 fmwr temperature (°c) june 27.17 ± 0.18 26.60 ± 0.06 27.78 ± 0.15 <35.0 6.5-8.5 july 26.67 ± 0.13 25.57 ± 0.19 27.17 ± 0.12 t-test, p-value 2.261, 0.087 5.316, 0.006 3.182, 0.033 ph june 7.10 ± 0.15 6.70 ± 0.12 6.50 ± 0.06 july 6.97 ± 0.18 6.93 ± 0.09 6.83 ± 0.09 t-test, p-value 0.571, 0.598 1.606, 0.184 3.182, 0.034 do (mg l–1) june 3.73 ± 0.15 5.67 ± 0.15 3.77 ± 0.09 7.5 july 3.40 ± 0.15 5.53 ± 0.15 3.93 ± 0.12 t-test, p-value 1.581,0.189 0.649, 0.552 1.118, 0.326 cod (mg l–1) june 25.17 ± 0.12 24.36 ± 0.10 26.67 ± 0.12 july 25.56 ± 0.13 24.41 ± 0.10 25.67 ± 0.04 t-test, p-value 2.226, 0.090 0.401, 0.709 7.973, 0.001 bod (mg l–1) june 4.38 ± 0.20 3.85 ± 0.10 4.74 ± 0.14 8.3 july 5.90 ± 0.06 4.05 ± 0.10 5.67 ± 0.07 t-test, p-value 7.233, 0.002 1.411, 0.231 5.899, 0.004 potassium (mg l–1) june 3.27 ± 0.17 2.58 ± 0.01 4.13 ± 0.05 july 4.16 ± 0.07 3.01 ± 0.06 4.30 ± 0.13 t-test, p-value 4.866, 0.008 6.697, 0.003 1.206, 0.294 magnesium (mg l–1) june 3.68 ± 0.16 3.13 ± 0.07 3.65 ± 0.05 july 4.29 ± 0.15 3.37 ± 0.09 4.64 ± 0.04 20 t-test, p-value 2.779, 0.050 2.040, 0.111 14.569, 0.0001 sodium (mg l–1) june 2.22 ± 0.15 2.38 ± 0.41 2.17 ± 0.07 200 july 1.71 ± 0.12 1.33 ± 0.06 2.51 ± 0.08 t-test, p-value 2.718, 0.053 2.539, 0.064 3.307, 0.030 calcium (mg l–1) june 2.48 ± 0.09 2.17 ± 0.17 2.80 ± 0.12 150 july 2.30 ± 0.15 1.93 ± 0.09 2.90 ± 0.12 t-test, p-value 1.026, 0.363 1.265, 0.275 0.612, 0.573 phosphate (mg l–1) june 1.01 ± 0.04 0.90 ± 0.05 1.16 ± 0.09 < 13.5 july 1.23 ± 0.13 0.88 ± 0.04 1.30 ± 0.12 t-test, p-value 1.568, 0.192 0.313, 0.770 0.938, 0.401 nitrate (mg l–1) june 3.90 ± 0.15 2.70 ± 0.06 4.90 ± 0.06 50 july 4.60 ± 0.21 2.43 ± 0.29 5.83 ± 0.09 t-test, p-value 2.711, 0.053 0.900, 0.419 8.854, 0.001 chloride (mg l–1) june 12.70 ± 0.12 10.53 ± 0.26 13.77 ± 0.19 250 july 13.53 ± 0.18 11.47 ± 0.03 14.00 ± 0.10 t-test, p-value 3.953, 0.017 3.556, 0.024 1.107, 0.330 sulphate (mg l–1) june 4.43 ± 0.23 3.50 ± 0.17 5.30 ± 0.06 100 july 5.20 ± 0.12 3.90 ± 0.10 5.57 ± 0.15 t-test, p-value 2.945, 0.042 2.000, 0.116 1.706, 0.163 fmev, nigerian federal ministry of environment; ph, hydrogen ion concentration; do, dissolved oxygen; cod, chemical oxygen demand; bod, biological oxygen demand. all values expressed as mean ± standard error mean (±sem). the level of significance between months was set at p<0.05. no nco mm er cia l u se on ly f.a. andong et al.78 successive fall in the level of dissolved oxygen concentrations due to mineralization of organic matter. such changes may retard the growth and reproduction of some fishes and in some severe conditions result to the death of the more vulnerable organisms. by the current ph records, lake oguta can be classified as weakly acidic to neutral, i.e. within the federal standard (6.5-8.5), and other records, e.g. 6-9 in (who, 2011); 6.5-8.9 (oluyemi et al.,2010) and 5.1-7.4 in (rim-rukeh, 2013). these results indicate a sufficient buffering property of the lake, suggesting safe agricultural and domestic uses (oluyemi et al., 2010). additionally, the current ph recorded may be adequate for the life of most aquatic organisms. in fact, a previous study recommended a range of 6.5-8.5 for most fishes to thrive (egemen, 2011). on the other hand, studies suggest that do value higher than 10 mg l–1 indicate bad or suboptimal conditions for the growth of aquatic fauna (clerk, 1986; bhatnagar and singh, 2010; ekubo and abowei, 2011). further, high do concentrations may indicate excessive algal proliferation (reynolds, 2006). here, the do recorded at the three stations was not beyond the federal limit (7.5 mg l–1), although quite low, and near the limit for most aquatic life (franklin, 2013). however, the do values for this study may be considered around or above values recommended for fish to survive (3-5 mg l–1; gorde and jadhav, 2013). the mean bod (3.9-5.9 mg l–1) was not beyond the federal limit (8.5 mg l–1) or who (2011) limits, and also not within other ranges from the same lake, e.g., 0.2-0.3 mg l–1 (nwoko et al., 2015). the cod, which is the amount of oxygen needed to carry out oxidation of organic waste by using strong oxidizing agent (awoyemi et al., 2014), was not beyond other standards, e.g. 2060 mg l–1 (who, 2011). however, the values were above other recorded values, e.g. 0.696.74 mg l–1 (oluyemi et al.,2010). phosphate level was below the federal standard (<13.5 mg l–1) and even below ranges reported in other studies (aduwo and adeniyi 2019). values above the federal standard may indicate pollution, because it is considered high (oecd, 1982). usually, sewage phosphate-based fertilizers used for agricultural activities are the cause of higher phosphate content in water. the nitrate content recorded here, are far below the federal (10 mg l–1) and the who (2011) standards (50.0 mg l–1), and below concentrations reported in a previous studies (37.2-43.9 mg l–1; igbinosa et al., 2012). higher phosphate and nitrate level promote eutrophication (ryding and rast, 1989). potassium concentrations (2.6-4.3 mg l–1) in the current study were slightly above other records (e.g. 2.12.6 mg l–1; aduwo and adeniyi, 2019), while sodium was far below the federal standard (200 mg l–1). sulphate and chloride values were below the fmwr standard. chloride can form many compounds (nacl, cacl2 and mgcl2), at varying concentrations in most natural waters (awoyemi et al., 2014). cloride is largely transported into the lake water by the dissolution of salts present in the soil and / or from polluting sources such as sewage and trade wastes (shaikh and mandre, 2009). higher values of chloride may affect water taste; however, these values are generally lower during the rainy than dry season (shaikh and mandre, 2009; awoyemi et al., 2014). other measured parameters like magnesium and calcium contribute to water hardness. total hardness of any water may be defined as the sum of calcium and magnesium concentrations and is normally expressed as milligrams of calcium carbonate equivalent per litre (karim and panda, 2014). in the present study, the hardness levels were very low compared to the federal standard and other studies (e.g. 229-1494 mg l–1; awoyemi et al., 2014). this is probably because there are less deposits of limestone materials around the lake, which is why the lake experienced low values of hardness (vermani and narula, 1995). conclusions the results presented in this paper indicates that some of the physical and chemical parameters in lake oguta were below nigerian federal ministry of water resources standard. for example, the physical parameters (temperature), the cations (calcium, magnesium and sodium), anions (phosphate, nitrate, chloride and sulphate) as well as bod and do were all below the fmwr standard. usually, the variations observed may be due to natural causes (geological, topographical, meteorological, hydrological and biological) and water level, as well as anthropogenic impacts. at present, no attempts were made to disentangle the contribution of natural and anthropogenic factors to the chemical conditions of the lake. further, considering the short period analysed, and to assure an accurate evaluation of the physical and chemical parameters, it is stressed that future studies should include a larger sample size and extended study periods (including all the seasons). acknowledgments the study was supported with facilities from the department of zoology and environmental biology and the laboratory of the national center for energy research and development, university of nigeria, nsukka, enugu state. we also thank the following people for their help with various aspects of this project: joseph effiong eyo, ike nelson ossai, okoye charles obinwanne and chinedu ifeanyi atama. no nco mm er cia l u se on ly water quality of lake oguta, nigeria 79 corresponding author: andongfelix@gmail.com key words: physical chemical variables; lake; nutrients; aquatic life; water quality assessment. conflict of interest: the authors declare no conflict of interest. received: 29 august 2019. accepted: 3 december 2019. this work is licensed under a creative commons attribution noncommercial 4.0 license (cc by-nc 4.0). ©copyright: the author(s), 2019 licensee pagepress, italy advances in oceanography and limnology, 2019; 10:8522 doi: 10.4081/aiol.2019.8522 references ademoroti cma, 1996. standard methods for water and effluents analysis. foludex press ltd., ibadan, nigeria. aduwo ai, adeniyi if, 2019. the physico-chemical water quality of the obafemi awolowo university teaching and research farm lake, ile-ife, southwest, nigeria. j. environ. prot. 10:881-899. doi: 10.4236/jep.2019.107052 ahiarakwem ca, nwankwor gi, onyekuru so, idoko ma, 2012. an assessment of the physical and environmental aspects of a tropical lake: a case study of the oguta lake watershed, niger delta basin, nigeria. am. int. j. contemp. res. 2:53-60. ahiarakwem ca, 2006. water quality assessment and some geochemical models of oguta lake, southeastern nigeria. phd thesis, federal university of technology owerri. ahiarakwem ca, onyekuru so, 2011. a comparative assessment of the physico-chemical and microbial trends in njaba river, niger delta basin, southeastern nigeria. j. water resour. prot. 3: 686-693. doi:10.4236/jwarp. 2011.39079. anitha g, 2002. hydrography in relation to benthic macroinvertebrates in mir alam lake hyderabad, andhra pradesh. phd thesis, osmania university, hyderabad. apha, 1998. standard methods for the examination of water and wastewater. apha, washington, dc. awoyemi om, achudume ac, okoya aa, 2014. the physicochemical quality of groundwater in relation to surface water pollution in majidun area of ikorodu, lagos state, nigeria. am. j. water. res. 2:126-133. doi: 10.12691/ajwr2-5-4. bhatnagar a, singh g, 2010. culture fisheries in village ponds: a multi-location study in haryana, india. agric. biol. j. n. am. 1:961-968. doi:10.5251/abjna.2010.1.5.961.968 clerk rb, 1986. marine pollution. clarendon press, oxford: 215 pp. dirican s, 2015. assessment of water quality using physicochemical parameters of çamlıgöze dam lake in sivas, turkey. ecologia 5:1-7. doi: 10.3923/ecologia.2015. 1.7. egemen o, 2011. water quality. ege university fisheries faculty publication, izmir 14:1-150. ekubo at, abowei jfn, 2011. review of some water quality management principles in culture fisheries. res. j. app. sci. eng. technol. 3:1342-1357. franklin pa, 2013. dissolved oxygen criteria for freshwater fish in new zealand: a revised approach. new zeal. j. mar. fresh. 48:112-126. doi.org/10.1080/00288330.2013.827123. garg rk, rao rj, saksena dn, 2009. water quality and conservation management of ramsagar reservoir, datia, madhya pradesh. j. environ. biol. 30:909-916. gorde sp, jadhav mv, 2013. assessment of water quality parameters: a review. int. j. eng. res. appl. 3:2029-2035. igbinosa eo, uyi oo, odjadjare ee, ajuzie cu, orhue po, adewole em, 2012. assessment of physicochemical qualities, heavy metal concentrations and bacterial pathogens in shanomi creek in the niger delta, nigeria. afr. j. environ. sci. technol. 6:419-424. doi: 10.5897/ ajest12.038. isinkaye mo, emelue hu, 2015. natural radioactivity measurements and evaluation of radiological hazards in sediment of oguta lake, south east nigeria. j. radiat. res. appl. sc. 48:1-12. doi: 10.1016/j.jrras.2015.05.001. karim aa, panda rb, 2014. assessment of water quality of subarnarekha river in balasore region, odisha, india. curr. world environ. 9:437-446. doi: 10.12944/cwe.9.2.27. mahananda hb, mahananda mr, mohanty bp, 2005. studies on the physico-chemical and biological parameters of a fresh water pond ecosystem as an indicator of water pollution. ecol. environ. conserv. 11:537-541. doi: 10.12691/ ajwr-2-5-4. nfor bn, akaegbobi im, 2012. inventory of the quaternary geology and the evolution of oguta lake, imo state, southeastern nigeria. world. j. pure app. sci. 22:56-53. nis, 2007. nigerian standard for drinking water quality: nigerian industrial standards, abuja 30 pp. nis, 2015. nigerian standard for drinking water quality: nigerian industrial standards. abuja nigeria 554:1-28. nwadiaro cs, 2018. fish introduction into lakes: a case for oguta lake, imo state, nigeria. futo journals 4:67-75. nwoko cia, ukiwe ln, egereonu uu, ukachukwu sn, 2015. assessment of seasonal physico-chemical parameters of oguta lake, nigeria. j. adv. chem. 11:3759-3764. doi: 10.24297/jac. v11i7.2199. odigi mi, nwadiaro cs, 1988. geophysical limnology of lake oguta (in imo state, south-eastern nigeria) with notes on its possible origin. hydrobiol. bull. 22:113-126. oecd, 1982. eutrophication of waters. monitoring, assessment and control. oecd, paris: 154 pp. okoro bc, uzoukwu ra, chimezie nm, 2014. river basins of imo state for sustainable water resources management. civ. environ. eng. 4:1-8. doi: 10.4172/2165-784x. 1000134. okorondu si, anyadoh-nwadike so, 2015. bacteriological and physiochemical analysis of oguta lake water, imo state nigeria. sc. j. public health. 3:14-19. doi: 10.11648/j. sjph.s.2015030501.13. oluyemi ea, adekunle as, adenuga aa, makinde wo, 2010. physico-chemical properties and heavy metal content of water sources in ife north local government area of osun state, nigeria. afr. j. environ. sci. technol. 4:691-697. ozoko dc, 2015. hydrochemistry of lakes in opi agu, enugu no nco mm er cia l u se on ly f.a. andong et al.80 state, southeastern nigeria. int. j. sci. technol. 4: 304-311. parashar c, dixit s, shrivastava r, 2006. seasonal variations in physico-chemical characteristics in upper lake of bhopal. asian. j. exp. sci. 20:297-302. patil pn, sawant dv, deshmukh rn, 2012. physico-chemical parameters for testing of water: a review. int. j. environ. sci. 3:11941207. reynolds cs, 2006. the ecology of phytoplankton. cambridge university press, cambridge: 507 pp. rim-rukeh a, 2013. physico-chemical and biological characteristics of stagnant surface water bodies (ponds and lakes) used for drinking and domestic purposes in niger delta, nigeria. j. environ. prot. sci. 4:920-928. doi: 10.4236/jep. 2013. 49106. ryding so, rast w, 1989. the control of eutrophication of lakes and reservoirs. unesco, paris: 314 pp. salla s, ghosh s, 2014. assessment of water quality parameters of lower lake, bhopal. appl. sci. res. 6:8-11. shaikh am, mandre pn, 2009. seasonal study of physicochemical parameters of drinking water in khed (lote) industrial area. int. res. j. 2:0974-2832. vermani op, narula ak, 1995. applied chemistry: theory and practice. new age int. publ. ltd., new delhi: 65 pp. who, 2011. guidelines for drinking-water quality. who, geneva: 541 pp. no nco mm er cia l u se on ly layout 1 introduction food web models provide a toolkit to study the relationship between species (or trophic groups) and communities. in order to better understand this relationship, structural (dunne et al., 2002) and dynamical (jordán et al., 2008) sensitivity analysis can be performed on food web models. the global (i.e., community-wise) effects or local (i.e. species-specific) disturbances can be simulated and quantified in several ways. these models can help to better understand several actual threats to aquatic ecosystems, including the appearance of introduced and invasive species, overfishing and local extinction of species. in all these cases, there is a major change concerning one focal species and effects spread out in the interaction network, influencing directly and indirectly many other species. in order to assess the effects on the whole community, we need general, predictive modelling tools. this way, we can better understand keystone species, indirect relationships, functional diversity and the vulnerability of communities. a central question is how do the effects of singlespecies perturbations (deletions) cascade across the network? the community effects of individual species may depend on their network position (pimm, 1980): this can be quantified by several network metrics (ranging from the number of neighbours to trophic height), quantifying how richly they are connected to others and where they are in the network. our key question is how to compare structural and dynamical predictions on the role individual species play in food webs. methods we consider here a food web simulation model, based on differential equations, where the community response is measured in various ways, following the disturbance of particular species. the model is as simple as possible, so we can focus on the pure differences between response functions, i.e. how to best measure the multi-species effects of single-species perturbations. our modelling framework is very general, applicable to ecological systems where trophic interactions dominate community dynamics (for example, open-water systems with a relatively low intensity of competition and non-trophic effects). generating networks and a dynamical simulation model we investigated 1000 randomly generated networks. by using some restrictions, they are similar and therefore comparable with a previously analysed network (móréh et al., 2018), which is based on a real food web (jordán et al., 2009; lin et al., 2004). the constraints and the common features were the following: all webs contain 15 trophic groups (3 producers and 12 consumers); there is constant number (36) of links between them; the number of top predators and trophic levels were maximized (maxtop ≤ 3, maxtl ≤ 4). this a simple model, suitable for dynamical analysis, representing basically a food web with low trophic resolution at lower levels and higher trophic resolution at higher levels. it remains for future studies to see how the aggregation procedure (i.e., describing microorganisms at higher details, see d’alelio et al., 2016, 2019; jordán et al., 2018) can possibly change the results. beyond these structural constraints our other criterion was the dynamical stability of the networks, and their robustness against perturbations. since the investigation of extinctions was out of our purposes, we looked for networks where all groups coexist during the whole article comparing community response indices in aquatic food web models ágnes móréh,1,2 ferenc jordán1,2,3* 1balaton limnological institute, centre for ecological research, tihany, hungary; 2evolutionary systems research group, centre for ecological research, tihany, hungary; 3stazione zoologica anton dohrn, naples, italy abstract aquatic ecosystems face several major challenges from the introduction and invasion of species, to overfishing. in order to better manage these situations, we need predictive models, where diverse scenarios can be simulated and tested. one key challenge to address is how to quantify the relationships between single-species disturbances and their multispecies effects. mapping the spread of direct and indirect effects in food webs helps to link species to communities. since food webs are complex networks of interactions, it is typically not easy to make predictions, so modelling and simulation may help to reveal general patterns. in food web simulations, one can quantify the effects of local perturbations on other species, i.e., community response. this may provide information about the relative importance of individual species and it is also useful to assess the vulnerability of the whole community to local changes. however, community response can be measured in several ways and various response functions give different results. in order to better understand their similarities and differences, we present a comparative study on a reasonable set of community response functions in food web simulations. these results contribute to build more predictive, multi-species models for systems-based conservation and management. no nco mm er cia l u se on ly aquatic food web models 95 simulation process, none of them extinct even in the course of perturbations. we modelled the dynamic behaviour of the networks in the same way as we did in móréh et al. (2018). the dynamics can be described as follows: (eq. 1) where bi means the biomass of species i, the consumption of the consumers can be characterised by a holling typeiii functional response (h = 2), where ωi is the relative consumption rate of species i when consuming ρ, b0 means the half-saturation density. the relative consumption rates are set to be proportional to the number of preys (ωi = 1/n). the increase function of the basal species (i ≤ 3) is described by the logistic growth model, where ri and ki are the intrinsic growth rate and the carrying capacity, respectively. the external mortality rates (di) of the basal species are set to 0. the increase and decrease of the consumer species (i > 3) depend only on the intensity of consumption of their predators and their preys (ri = ki = 0), and the external mortality rate, which is set to 0.1 for all consumers. since our purpose was to focus on the impact of network topology on dynamics, so we didn’t model the consumption and conversion rates explicitly, but assumed the strength of a predator-prey link (ε) proportional to the number of preys (ε = 1/n). for the sake of the same cause, the parameters (even the mortality rates) were fixed, only the different distributions of the links between the groups determinate the topology, thus the dynamics, too. the process of dynamical simulations had 3 main parts. first, the integration of the system (hindmarsh et al., 2005) until it settled to a fixed point; if any of the species is extinct, the integration was terminated and the adjacency-matrix was casted off. second, if all species reached a fixed point and coexisted, we made a preliminary stability analysis on the system by arbitrary changing the biomass-values of all species. after the system returned to the original equilibrium after this pulse perturbation event, we made a systematic perturbation process on all consumers (the producers were part of the dynamical system but their community effects were not evaluated). we changed the mortality rate (di) of the species in question increasing it by 10%. this analysis can be considered as a press perturbation experiment (sensu bender et al. (1984)). third, if the system is robust against all perturbation events (there is no extinction), not only the adjacencymatrix of the network, but also the biomass-values registered before and after all perturbation events are hold and stored. community response indices we are interested in the effect of perturbing species i on the ecosystem including it, thus we have to make a relationship between the original and the perturbed state of the system. however, only one species is perturbed, all others’ state changes; the whole system is settled to a new equilibrium state. in our modelling framework, we consider the biomass of species but also abundance could have been used (as another, measurable, quantitative trait). all species “answer” to the perturbation event by its own biomass-change; thus, the community response (cr) to the perturbation of the ith species (cri) can be determined as the sum of all these answers. there is not a single formula used to determine the response of community to any kind of perturbations, but more different approaches exist and are used in parallel. the investigated types and their variants are summarized in tab. 1. “functional importance index” (fii) is defined by hurlbert (1997), and calculates the summation of the differences of biomass-values measured before and after perturbation. this fact discriminates this index from all others discussing below, since the difference gives information about the net changes in a numerical way, but gives less information about the order of magnitude of the changes. all indices discussed further calculate the rate of biomasses, which, however, shows the measure of the changes, not the exact numerical values of them. “interaction strength index” (isii) used by okey (2004) calculates simply the sum of the rates of biomassvalues after and before perturbation. thus, this index takes not only the measure of the changes into account, but inherently the direction of the changes, too. “paine’s interaction strength” (pii, (paine, 1992)) can be derived from okey’s interaction strength index by subtracting 1 from the rates calculated in isii. in fact, these indices are complements of each other: while the isii gives the exact rate of the biomass-values, the pii shows the measure of changes. in addition, pii enables to ignore the sign of changes by calculating absolute values (tab. 1). the next two indices incorporate a further aspect, namely the concept of keystone species considering its core meaning: “a keystone species is one whose effect is large, and disproportionally large relative to its abundance” (power et al., 1996). the index named “keystoneness” (kni) is also derived from isii divided it by the percent of the system’s overall living biomass represented by group i before it was perturbed (okey, 2004). the “community importance index” (cii) also provides an index of the impact of a species relative to its biomass no nco mm er cia l u se on ly á. móréh, and f. jordán96 (power et al., 1996). it calculates the sum of the opposite (sign) of the real value of the change of all species divided by the percent of the overall biomass of the groups represented by species i before it was perturbed. although all discussed indices are based on the comparison of the biomass-values before and after the perturbation event, and the summation of the changes of each groups, there are two important points of view, that discriminate the four variants of them (columns in tab. 1): i) taking the sign of the biomass-changes of every single species into account or not; ii) taking the self-effect of the perturbed species into account or not (this means feedback: if a particular species is disturbed, effects spread in the network and indirect chain effects can influence the disturbed species later again, looping back to it). an additional distinction between the different types of crindices is, whether they incorporate the original biomass of the perturbed species (considering the keystoneconcept), or not. structural network indices in order to quantify the position of individual species in food webs, we calculated 18 structural importance indices that are summarized in tab. 2. for detailed description of the indices see móréh et al. (2018). although all of the 18 indices express the positional importance in some wise, they differ from each other, too. there are local indices not taking indirect effects into account (e.g., d), while there are meso-scale or non-local indices (e.g., centralities). the topological importance indices consider binary interactions, others (e.g., wi) can quantify weighted webs. while some of them (e.g., s, k) characterise directed networks, others do undirected ones (e.g., d). correlation analysis we used the non-parametric spearman’s rank correlation test to investigate the connection between the different cr-functions. beyond that we studied the correlations of crs both with the measures of self-effects and the different structural network indices. its purpose was to investigate the consistency of the relationships between the indices and the different types and variants of crs. since we were interested in the strength of correlations, not the sign of them, we used the absolute values of the ρ-values for further analysis. we considered a connection strong if |ρ| ≥ 0.5. results correlation between the different response functions in tab. 3, we summarize not only the strengths of correlations (|ρ|) between the different response functions, but also between the measure of self-effects or original biomass-value of the perturbed species and the cri-types. we considered the correlation stronger if |ρ| ≥ 0.5. investigating the relationships between the different variants within the same type we can see that the indices calculated with or without self-effect (in grey in tab. 3) tab. 1. summary of all investigated indices: the main types re shown in rows, while the different variants are collected in columns. bibef and biaft denote the biomass values before and after the perturbation event; %bibef is the percentage of the affecting consumer in the whole system before it was perturbed (%bbef = (bi/σn1bj)x100,n = 15). the original form of each types used in the literature are highlighted with grey background; variants calculated with self-effect (i ∈ n, where n = {1...15}) are marked with star. sign ignored with sign references i ∉ n i ∈ n i ∉ n i ∈ n functional importance index (fii) |fii|* fii* hurlbert (1997) livi et al. (2011) interaction strength index (isii) isii* okey (2004) paine’s interaction strength index (pii) |pii|* pii* paine (1992) móréh et al. (2018) keystoneness (kni) kni* okey (2004) community importance index (cii) |cii|* cii* power et al. (1996) okey (2004) no nco mm er cia l u se on ly aquatic food web models 97 show stronger correlations in most cases (|fii| vs. |fii|* shows the strongest, fii vs. fii* shows the weakest correlation). in turn, indices taking the sign of changes into account or ignoring them are much less correlated (framed values in tab. 3). investigating the connections between the different types tab. 2. summary of the structural network indices. index name notation reference degree d wassermann and faust (1994) weighted degree wd centrality indices wassermann and faust (1994) betweenness centrality bc closeness centrality cc keystone index and its components jordán et al. (1999) keystone index k bottom-up effects kbu top-down effects ktd direct effects kdir indirect effects kindir status index and its components harary (1959) status s contra-status s’ delta-status δs positional importance based on indirect effects jordán et al. (2003) topological importance, max step = 1 ti1 topological importance, max step = 3 ti3 topological importance, max step = 5 ti5 weighted topological importance max step = 1 wi1 weighted topological importance max step = 3 wi3 weighted topological importance max step = 5 wi5 tab. 3. the strengths of spearman-correlations (|ρ|) between the different cr-types and variants. bold values mean higher degrees of correlations (|ρ| ≥ 0.5). relationships between the variants calculated with or without self-effect within the same type are marked with grey background. values of correlations between indices taking the sign of changes into account or ignoring them (within the same type) are framed. |fii| |fii|* fii fii* isii isii* |pii| |pii|* pii pii* kni kni* cii cii* |cii| |cii|* |fii| 1 0.98 0.32 0.47 0.31 0.16 0.93 0.84 0.31 0.16 0.65 0.66 0.23 0.36 0.39 0.22 |fii|* 1 0.43 0.46 0.38 0.14 0.93 0.89 0.38 0.14 0.65 0.66 0.3 0.34 0.4 0.26 fii 1 0.35 0.58 0.06 0.36 0.58 0.58 0.06 0.1 0.09 0.55 0.01 0.6 0.66 fii* 1 0.07 0.14 0.45 0.27 0.07 0.14 0.77 0.76 0.17 0.36 0.33 0.44 isii 1 0.67 0.27 0.39 1.0 0.67 0.09 0.1 0.94 0.57 0.26 0.3 isii* 1 0.12 0.04 0.67 1.0 0.32 0.32 0.53 0.92 0.23 0.31 |pii| 1 0.9 0.27 0.12 0.67 0.67 0.18 0.34 0.46 0.26 |pii|* 1 0.39 0.04 0.44 0.45 0.36 0.12 0.64 0.54 pii 1 0.67 0.09 0.1 0.94 0.57 0.26 0.30 pii* 1 0.32 0.32 0.53 0.92 0.23 0.31 kni 1 1.0 0.03 0.59 0.31 0.46 kni* 1 0.02 0.59 0.3 0.46 cii 1 0.44 0.32 0.4 cii* 1 0.28 0.41 |cii| 1 0.94 |cii|* 1 self-effects |diffi| 0.74 0.84 0.68 0.28 0.5 0.06 0.74 0.86 0.5 0.06 0.46 0.47 0.45 0.19 0.42 0.41 diffi 0.74 0.84 0.68 0.28 0.5 0.06 0.74 0.86 0.5 0.06 0.46 0.47 0.45 0.19 0.42 0.41 ratei 0.42 0.53 0.84 0.16 0.53 0.14 0.41 0.71 0.53 0.14 0.09 0.08 0.57 0.16 0.68 0.79 |ratei-1| 0.42 0.53 0.84 0.16 0.53 0.14 0.41 0.71 0.53 0.14 0.09 0.08 0.57 0.16 0.68 0.79 ratei-1 0.42 0.53 0.84 0.16 0.53 0.14 0.41 0.71 0.53 0.14 0.09 0.08 0.57 0.16 0.68 0.79 %bibef 0.66 0.66 0.09 0.76 0.11 0.33 0.67 0.45 0.11 0.33 1.0 1.0 0.01 0.6 0.3 0.46 no nco mm er cia l u se on ly á. móréh, and f. jordán98 and variants of cris (indices possibly considering sign and or self-effect), we can see that the correlation can be stronger if the compared indices take sign into account inherently (isii, kni) or are calculated without absolute value (pii, cii). in case of indices can be calculated with or without sign (fii, pii, cii), the correlation is stronger if sign is ignored (e.g., |fii| vs. |pii|) regardless of ignoring self-effects or not. the question is whether an index uses the difference or the rate of biomass-values measured before and after the perturbation event. fii-indices use the difference of these, thus, it is strongly dependent on the original values of not only the perturbed (bi), but also the all other species’ biomass. fii (calculated with sign and with selfeffect) is and exception: its correlation is very weak with the original percentage of the perturbed species’ biomass. using the rate of the biomasses makes the changes independent of the real biomass-values, these indices emphasizes the measure of the changes. |fii|* and |pii|* show the strongest correlations with any types of selfeffects, while the cii-type indices correlate stronger only with the rate-types of self-effects (excepting cii*, which is, however, in stronger correlation with the original percentage of the perturbed species’ biomass). while isii shows a moderate correlation with the strength of selfeffects, dividing it by %bibef deteriorates this connection by including the dependency of the original amount of the perturbed species (kni). the difference between the variants within the same type can be enlightened also by comparing the 12,000 original values of each types (1000 networks, 12 perturbation events). fig. 1a shows the difference of the indices calculated with and without self-effect. the largest impact of self-effect can be seen in cases of kni and isii, while the cii-indices show smaller difference in this regard. the variants with self-effect are almost always greater than the variants ignoring it, except fii and pii, where this difference is always negative (marked with grey boxes in fig. 1). investigating the differences between the variants calculated with or without taking sign into account, we get that the variants without sign are always greater than the variants calculated with absolute value (fig. 1). it means that using sign always underestimates the strength of perturbation effects because of the changes with opposite sign cancelling each other out. correlation between the different crand structural network indices the results of the comparison of the different crivalues and the structural indices of each investigated networks are shown in fig. 2. first of all, there is a conspicuous difference between two groups of network indices. while the centrality indices (cc, bc), the unweighted versions of topological importance indices (tin), the keystone index (k) and its components referring the direct or indirect effects for a node (kdir, kindir) show very weak connection with almost all types of cr-indices, the status-indices (s, s’, ∆s), the weighted versions of topological importance indices (win) and the degree (wd), and the components of keystone index quantifying the bottom-up and the top-down effects of a node (kbu, ktd) show higher average correlation strength; in case of some particular indices the connection is quite strong. examining the different response-indices more carefully we can see that the types ignoring not only the sign of changes but also the original percentage of the perturbed species (|fii|, |fii|*, |pii|, |pii|*) show the strongest correlations with all network indices belonging the group that show higher average correlation with the response-indices. in this regard, the self-effect of the fig. 1. the results of the comparison of the variants of the different cr-types. a) difference of types calculated with self-effect or ignoring it; the boxplots show the absolute values of the difference; the originally negative values are marked by grey colour. b) difference of variants calculated with or without sign. all values are positive. note, that kni and isii have been calculated with sign. no nco mm er cia l u se on ly aquatic food web models 99 perturbed species does not matter. taking sign into account the strength of correlations is rather moderate, or weak in some cases, and the differences also are much smaller. moreover, ignoring not only the keystone concept (%bbef) but also the self-effects causes a bit stronger correlation. when sign is ignored, calculating with %bbef (|cii| and |cii|*) can increase the correlation values to a bit higher level in case of some network indices referring also to some aspects of keystoneness (e.g. k, kdir), but these strengths remain still rather weak. in contrast with that, in case of status indices the strength of correlations is much higher, albeit not as strong as ignoring %bbef. all in all, in case of the structural indices showing stronger correlations, the strength is influenced much more by sign than self-effect. however, the difference of variants with or without self-effect is not as unambiguous as in case of the sign, the averages of correlation-values in cases where self-effects are ignored are higher. discussion considering the powers and the limits of these different formulas of community response functions we can choose the one that suits our purposes best. if our study is confined only the investigation of a single food web, in respect of the way of comparison it does not matter whether we use a formula using the difference or the rate of the biomass-values. although neither the exact numerical difference nor the rate of change can handle the two important point of view of a change together: a large difference doesn’t mean large rate of change if the original biomass is high enough, in the same way, small difference can give both small and large rate depending on the original values. despite that in case of studying more networks it might be more useful comparing only the magnitude of changes between the different networks increasing the comparability of the results. likewise, the inclusion of self-effect in the sum is fig. 2. the strength of correlation (|ρ|) between the studied network indices and the different cr-types grouped into columns based on the three investigated point of view, whether the sign of changes, the self-effect of the perturbed species and the keystone concept are considered or not. the network indices (rows) are sorted by descending the average correlation strength between all cr-indices and the structural index in question. no nco mm er cia l u se on ly á. móréh, and f. jordán100 determined by the current question. in case of rate-type cr-functions the correlations between the two variants are very good. in case of functional importance indices there can be larger differences between the results calculated by the two ways because of the higher dependency of the original biomass-values of the perturbed species (that is supported by the higher correlations with the percentage of it; tab. 3). if the investigation of the connections between the species’ structural positions and any changes in a trait of them (say the biomass) is among the purposes of a study, including or ignoring self-effects has no particular impact on the strength of correlations; although the values without self-effects are higher in general, the differences are very small. all in all, from this point of view, the question of ignoring sign is much more important than ignoring selfeffect, or even more calculating the difference or the rate of biomasses. calculating with sign underestimates the net measure of changes because of changes with opposite sign can cancel out each other. theoretically, it is possible that while a corresponding cr-index gives the result of 0, the whole structure of the ecosystem changes. since there are indices, that inherently take sign into account (isii, kni), it is better to use a form with an equivalent “structure” (e.g., pii instead of isii). this problem can also lead up to different conservation biology approaches. comparing small negative and large positive effects, we can understand if it is better to minimize our impact on nature (small negative better than large positive) or to implement positive effects (large positive better than small negative). choosing a response function that does not provide information about increase or decrease of population size, only about the change in population size is supported by a conservation philosophy suggesting that minimizing the human impact (size of change) might be preferred over trying to help natural systems (direction of change). conclusions the community-wide effects of single-species perturbations can be modelled, simulated and measured. mapping the spread of effects in food webs is possible from several approaches, including a general bottom-up vs top-down comparison (stäbler et al. 2019), in singlenode vs pairwise perturbation simulation studies (móréh et al. 2018) or in single-node vs multi-node structural analyses (jordán et al. 2019). linking individual species to communities may also consider reproductive systems (d’alelio 2017), trait-based aggregation protocols (jordán et al. 2018) or the relationship between carbon fluxes and interaction strengths (scotti et al. 2012). better databases, including long-term ecological data (d’alelio et al. 2016) provide a robust empirical basis for ecosystem modelling (d’alelio et al. 2019) and all these approaches may contribute to multi-species, system-based conservation efforts and fisheries management. some results presented in this paper contributes to more predictive modelling. based on tab. 3, we see, for example, that simulating the effects of overfishing does not need to consider self-effects (feedback) if the sign (good or bad effect) of the interaction is not considered, only its magnitude (big or small effect). if signs are considered, however, calculating self-effects can be important for predictability. fig. 2 suggests that the best structural predictions can be achieved when neither effect sign nor initial biomass are considered. in this case, weighted and possibly also indirect structural indices are the best ones to predict systems dynamics. these hints help to offer predictive simulation models for fisheries in aquatic ecosystems. acknowledgements we thank sándor imre piross and anett endrédi for useful advices. the research was supported by the national research, development and innovation office (nkfih) grants otka k 116071 and ginop-2.3.2-152016-00057. corresponding author: jordan.ferenc@gmail.com key words: food web; keystone species; perturbation; community response. received: 16 october 2019. accepted: 23 december 2019. this work is licensed under a creative commons attribution noncommercial 4.0 license (cc by-nc 4.0). ©copyright: the author(s), 2019 licensee pagepress, italy advances in oceanography and limnology, 2019; 10:8621 doi: 10.4081/aiol.2019.8621 references bender ea, case tj, gilpin me, 1984. perturbation experiments in community ecology: theory and practice. ecology 65:1-13. d’alelio d, libralato s, wyatt t, ribera d’alcalà m. 2016. ecological-network models link diversity, structure and function in the plankton food-web. sci. rep. 6:21806. d’alelio d, hay mele b, libralato s, ribera d’alcalà m, jordán f. 2019. rewiring and indirect effects underpin modularity reshuffling in a marine food web under environmental shifts. ecol. evol. 9:1-16. no nco mm er cia l u se on ly aquatic food web models 101 dunne ja, williams rj, martinez nd, 2002. network structure and biodiversity loss in food webs: robustness increases with connectance. ecol. lett. 5:558-567. harary f, 1959. status and contrastatus. sociometry 22:23. hindmarsh ac, brown pn, grant ke, lee ke, serban r, shumaker de, woodward cs, 2005. sundials: suite of nonlinear and differential/algebraic equation solvers. acm trans. math. softw. 31:363-396. hurlbert sh, 1997. functional importance vs. keystoneness: reformulating some questions in theoretical biocenology. austral ecol. 22:369-382. jordán f, takács-sánta a, molnár i, 1999. a reliability theoretical quest for keystones. oikos 86:453-462. jordán f, liu w, van veen f, 2003. quantifying the importance of species and their interactions in a host-parasitoid community. comm. ecol. 4:79-88. jordán f, okey ta, bauer b, libralato s, 2008. identifíing important species: linking structure and function in ecological networks. ecol. mod. 216:75-80. jordán f, liu w, mike á, 2009. trophic field overlap: a new approach to quantify keystone species. ecol. mod. 220:2899-2907. jordán f, endrédi a, liu wc, d’alelio d. 2018. aggregating a plankton food web: mathematical versus biological approaches. mathematics 6:336. jordán f, pereira j, ortiz m. 2019. mesoscale network properties in ecological system models. curr. opin. syst. biol. 13:122128. lin hj, shao kt, hwang js, lo wt, cheng ij, lee lh, 2004. a trophic model for kuosheng bay in northern taiwan. j. mar. sci. technol. 12:424-432. livi cm, jordán f, lecca p, okey ta, 2011. identifying key species in ecosystems with stochastic sensitivity analysis. ecol. mod. 222:2542-2551. móréh á, endrédi a, jordán f, 2018. additivity of pairwise perturbations in food webs: topological effects. j. theor. biol. 448:112-121. okey ta, 2004. shifted community states in four marine ecosystems: some potential mechanisms. phd thesis, university of british columbia, vancouver. paine rt, 1992. food-web analysis through field measurement of per capita interaction strength. nature 355:73-75. pimm sl, 1980. food web design and the effect of species deletion. oikos 35:139-149. power me, tilman d, estes ja, menge ba, bond wj, mills ls, daily g, castilla jc, lubchenco j, paine rt, 1996. challenges in the quest for keystones. bioscience 46:609-620. scotti m, gjata n, livi cm, jordán f. 2012. dynamical effects of weak trophic interactions in a stochastic food web simulation. community ecol. 13:230-237. stäbler m, kempf a, smout s, temming a, 2019. sensitivity of multispecies maximum sustainable yields to trends in the top (marine mammals) and bottom (primary production) compartments of the southern north sea foodweb. plos one 14:e0210882. wassermann s, faust k, 1994. social network analysis: methods and applications. cambridge university press, cambridge: 825 pp. no nco mm er cia l u se on ly layout 1 introduction any submerged surface exposed to light is likely to be colonized by more or less adherent benthic phototrophic organisms. the carapace of turtles living in aquatic environments is no exception. herpetologists and algologists have long been interested in algal species living as epibionts on freshwater and marine turtles (basibionts) (edgren et al., 1953). the relationship between turtle and algae is considered commensal, since algae take advantage of the movement of their host for access to resources and dispersion, while algae generally have little effect on the health of their host (edgren et al., 1953; neil and allen, 1954; wahl, 2008). it can even be seen as a kind of mutualism because algae can provide camouflage for turtles against predators. the study of epibiosis in turtles has historically focused on macro-epibionts and especially macroalgae (neil and allen, 1954; frick and pfaller, 2013), while epizoic microalgae have only recently been investigated (wu and bergey, 2017). diatoms, because of their high adhesion capacity and rapid growth, are well-known early colonizers of submerged surfaces (wahl, 1989; biggs et al., 1998). because of their ubiquitous nature and high specific diversity, benthic forms are widely used as biological indicators of water quality (poulíčková and manoylov, 2019). the last decade has seen a growing interest in diatoms on turtle shells (wetzel et al., 2010; majewska et al., 2015; wu and bergey, 2017; donato et al., 2018; vassal et al., 2020). it is possible to distinguish species-level studies focusing on the search for new taxa, probably specific to epizoic habitat (wetzel et al., 2012; majewska et al., 2020) and others at the level of diatom communities, focusing on variations observed at the host individual level, or between turtle populations or species (fayolle et al., 2016; robinson et al., 2016; wu and bergey, 2017). the purpose of the latter is sometimes to obtain information on the environments frequented by individuals (vassal et al., 2020) and on their movements. such information can be valuable considering that many turtle species are in danger of extinction (rhodin et al., 2018). however, the discovery of new species supposedly restricted to the epizoic habitat and the tracing of turtle movements by studying the communities present on their carapace are based on different hypotheses concerning the selectivity of the carapace substrate. in the first case, the carapace of the turtles is assumed to be highly selective so that new taxa discovered in this habitat are considered typically epizoic, thus forming part of the obligatory epibionts. in the second case, the epizoic flora is supposed to reflect the surrounding benthic communities in the turtle’s living environment, with diatoms being seen as facultative epibionts (wahl and mark, 1999) and the turtle carapace as a neutral substrate (blindow, 1987). in order to be able to determine whether turtle carapace is a neutral substrate or whether it selects a specific community, it is necessary to sample, at each study site, several host individuals and other types of substrates, such as stones or macrophytes, generally used for biofilm sampling in running waters and lakes. this has generally not been done formally in the above-mentioned studies on turtle-diatom relationships and yet it is the only way to characterize the specificity of diatom communities living on turtle carapaces. the european pond turtle emys orbicularis (linnaeus article specificity of diatom communities attached on the carapace of the european pond turtle (emys orbicularis) vincent roubeix,1* lucie attia,2 rémy chavaux,3 franck very,4 anthony olivier,5 luc ector,6 véronique vassal7 1hydrobiostat, aix-en-provence, france; 2dreal occitanie, laboratoire d’hydrobiologie, cité administrative bât g, 1 rue de la cité administrative, cs 8002, 31074 toulouse cedex 9, france; 3dreal auvergne-rhône-alpes, laboratoire d'hydrobiologie, 5 place jules ferry, 69453 lyon cedex 06, france; 4dreal auvergne-rhône-alpes, laboratoire d'hydrobiologie, 7 rue léo lagrange, 63000 clermont-ferrand, france; 5la tour du valat, institut de recherche pour la conservation des zones humides méditerranéennes, le sambuc, 13200 arles, france; 6environmental research and innovation (erin) department, luxembourg institute of science and technology (list), 41 rue du brill, 4422 belvaux, luxembourg; 7dreal paca, laboratoire d’hydrobiologie, pôle d’activités les milles, avenue albert einstein, bâtiment e cerema, dter med 13593m, cs 70499, 13593 aix-en-provence cedex 3, france abstract periphytic diatoms growing on aquatic turtle carapaces provide an interesting example of biofouling on an animal substrate. this study dealt with the question of the specificity of epizoic diatom communities for the case of the european pond turtle (emys orbicularis), across five water bodies in south-eastern france; 375 diatom taxa have been identified on the european pond turtle carapace. the results showed that the epizoic communities of the five sites were significantly different from each other. the epizoic diatom communities could be clearly distinguished from epilithic and epiphytic communities only in two out of five sites. the differences in composition at the two sites resulted from the greater abundance of some species and life forms. however, the characteristic pattern of diatom communities on the carapaces was not the same at the two sites. it therefore appeared that the european pond turtle’s carapace did not harbour specific diatom communities and can rather be considered a neutral substrate. v. roubeix et al.2 1758) is a medium-sized freshwater turtle with a wide ecological range. its home range size is limited by physical barriers which means that individuals can use the whole surface of a wetland with a preference for habitats with a strong plant density (cadi et al., 2008). during its period of activity (generally from march to october), it comes out of the water for its thermoregulation needs (di trani and zuffi, 1997) and for egg-laying by females (rovero and chelazzi, 1996). more rarely, individuals can migrate on land or estivate during droughts and low-water periods (naulleau, 1992). due to its significant decline, e. orbicularis benefits from a national action plan for protection in france (thienpont, 2020) and is listed in annexes ii and iv of the eu habitats directive. diatom species have already been recorded on the european pond turtle in turkey, france and serbia (soylu et al., 2006; akgul et al., 2014; ersanli and gonulol, 2014; fayolle et al., 2016; vidakovic et al., 2018), but the specificity of carapace communities in relation to other natural substrates has only recently been addressed (vassal et al., 2020). the suitability of diatoms collected on turtle carapace for water quality bioindication was assessed in two mediterranean sites (vassal et al., 2020). the investigation was extended in the present study by the inclusion of three other sites and a closer analysis of the specificity of diatom epizoic communities. periphytic diatoms were sampled on three different types of substrates: on european pond turtle carapace, on stones and on macrophytes. for practical reasons related to the analysis of the data, the question of the specificity of the epizoic communities was divided into three sub-questions: i) are the diatom communities of the european pond turtle carapace the same among sites? this is the question generally raised in large-scale epizoic community studies (wu and bergey, 2017) and a similarity of communities sampled at different sites may suggest that the epizoic flora is specific; ii) are there typical species of the epizoic habitat? this is the question that generally interests taxonomists and implies a selection of specific taxa more adapted to the habitat, as a result of a co-evolution of host and epibiont species; iii) at each site, what are the differences between epizoic communities and those of other substrates? this third sub-question investigates whether epizoic diatom communities are representative of the periphytic communities at a given site and similar to those commonly sampled on stones or plants. substrate neutrality implies no significant species selection, but does not preclude variation in community density, which depends on various factors related to the host and its environment (frick and pfaller, 2013). for this reason, the study focused only on community composition and relative abundance. the investigation was carried out through two scales of space (inter-sites and intra-site) and two levels of biological organization (community and species). methods periphytic biofilm sampling periphytic diatoms were sampled in five french water bodies of various types, hereafter identified by their code (ver, sav, esq, cha, tar), where populations of the european pond turtle are present and monitored (fig. 1, tab. 1). all sampling sites are small or medium lentic systems: one is brackish (tar) and another is eutrophic with low o2 concentration (cha). at sites cha, esq and tar, where no mineral substrate was available, pebbles collected on the shore of the lakes were placed in a permanently submerged area at least one month prior to sampling. at each site epilithic biofilm was collected from at least five pebbles and mixed to form an integrated sample. similarly, epiphytic biofilm was collected from at least five stems of helophytes present in the water body and mixed. the sampling of epilithic and epiphytic diatoms followed the national protocol defined for bioindication in water bodies (irstea, 2013). the turtles were captured by hand or with fyke nets. the epizoic biofilm was obtained by brushing with a toothbrush an area of 20 cm² on the back of the carapace of ten adult turtle individuals (only seven for the ver site where no more individuals could be captured) (fig. 2). the biofilm on the carapaces was generally very thin, even invisible, and no macrophyte was apparent on the sampled area. collection from each individual provided one epizoic sample. analytical effort was concentrated on the animal substrate that is least known, so that site-specific replicates were considered only for epizoic communities. in order to facilitate comparison between substrates within a site, an additional integrated sample was prepared by mixing subsamples of all singular epizoic samples. thus, the number of samples per site was 13 in total (10 for site ver). for each site, the sampling of the three substrate types was done within 24 h, in spring or early summer (tab. 1), assuming little temporal variations in differences of periphytic communities among substrate types over the months of sampling. samples were fixed in 90% ethanol until preparation for analysis. sample preparation and diatom counts the samples were digested in hot hydrogen peroxide (90°c) for several hours and then treated with concentrated hydrochloric acid. after rinsing and decanting, they were mounted on microscope slides and fixed with naphrax® before being observed under an optical microscope. at least 400 diatom valves were counted in each sample. the species were identified according to the specific literature (lange-bertalot et al., 2017). the determination of certain difficult taxa required examination with a scanning electron microscope. diatoms on the european pond turtle carapace 3 data analysis the analysis of abundance data of diatom species in periphytic communities was carried out using r software and the vegan and labdsv packages (oksanen et al., 2019; roberts, 2019). only species with a maximal relative abundance exceeding 1% were considered in the statistical treatments on communities. as species abundance in communities do not comply with the hypothesis of multinormality, standard parametric tests could not be applied. therefore, the similarity of epizoic communities between sites was tested using nonparametric multivariate analysis of variance (np manova, anderson, 2001) with canberra distance on untransformed relative abundance data (n=47, 9999 permutations). the homogeneity of community variance between sites was previously tested and found to be acceptable (anova; p=0.16). the samples were represented on a plot by principal coordinate analysis (pcoa) using the same canberra distance. the np manova was then repeated ten times to compare epizoic communities between each pair fig. 1. location of the five study sites. table 1. synthetic data about the five study sites (nd = not determined). code name location geographic type of water sampling macrophytes temp. ph cond. o2 coordinates body date used for (°c) (µs/cm) (mg/l) sampling ver verrerie thionne 46.419450°n shallow lake 06/21/2018 juncus sp. 27.8 6.9 30 8.6 3.561080°e sav save arandon-passins 45.703177°n lake 04/24/2018 carex sp. 19.9 7.7 564 9.1 5.452998°e esq esquineau arles 43.509970°n marshes 06/08/2017 phragmites 26.2 7.8 459 6.47 4.655926°e australis cha charpines la roque d’anthéron 43.734896 °n oxbow of 06/29/2016 phragmites 26 nd 1360 4 5.313876 °e the durance australis river tar tartuguière lansargues 43.628857°n brackish 05/15/2018 phragmites 13.8 8 9110 nd 4.094122°e marshes australis v. roubeix et al.4 of sites. shannon diversity and species richness of samples were compared among sites and type of substrate (animal vs non animal) in each site using the non-parametric kruskal-wallis and wilcoxon tests, respectively. in order to identify taxa representative of the epizoic habitat, the indicator value (indval) method (dufrêne and legendre, 1997) was used on all data, without abundance threshold. the analysis first focused on all the sites. the group considered was the mixed epizoic samples (n=5), compared to epilithic (n=5) and epiphytic (n=5) samples. then, the analysis was applied to each site, considering all the individual epizoic samples (n=7 or 10), compared to the two non-epizoic samples (n=2). differences in the composition of all samples in the study (n=62) were analysed by ascending hierarchical classification using ward’s method and canberra distance. the resulting dendrogram was cut at two distance levels to highlight relevant groups. at sites where differences between substrates were apparent, the epilithic, epiphytic and mixed epizoic samples were compared on the basis of their composition into functional groups. the assignment of functional groups to species or genera was done according to the classification of taxa proposed by rimet and bouchez (2012). chi-square tests were carried out to assess in each site and among substrate types (1) the similarity of functional composition of diatom communities and (2) the equality of proportions of some life forms. results diatoms were found on the carapaces of all individuals of emys orbicularis used in the study (n=47). a total of 375 taxa were identified in the epizoic samples. the absence of macrophytes on the carapaces suggested that these diatoms were directly epizoic and not epiphytic and secondarily epizoic. the diversity and species richness of diatom samples varied significantly among sites (p<10-6). there was a significant effect of the type of substrate only fig. 2. sampling of epizoic diatoms on the carapace of an individual of emys orbicularis. fig. 3. dendrogram derived from the ascending hierarchical classification of the 62 diatom samples and barplot of species number with indication of shannon diversity. the codes ‘lith’ and ‘phy’ mean epilithic and epiphytic. the code ‘zmi’ corresponds to the mixed epizoic sample whereas the codes ‘z01’ to ‘z10’ are for the individual epizoic samples. stars, diamonds and triangles indicate mixed epizoic, epilithic and epiphytic samples, respectively. diatoms on the european pond turtle carapace 5 on species richness and at two sites (p≤0.05). indeed, epizoic samples were richer in species than non-epizoic samples in the sites ver and esq (fig. 3). different epizoic communities among study sites considering the dendrogram resulting from the ascending hierarchical classification, five groups could be clearly distinguished, defined by a cutting level at a distance of 300 (fig. 3). these five groups corresponded exactly to the five study sites, meaning that the periphytic communities depended more on the water body they came from, than on the type of substrate from which they were extracted. the species richness of the samples was highest at the esq and sav sites (40 to 50 species, fig. 3). the non-parametric multivariate analysis of variance indicated that the epizoic communities of the five sites were significantly different (p=10-4) and that the communities at each site differed significantly from those at each of the other sites (p≤2 10-4). these differences between the sites were clearly visible in the principal coordinate analysis (fig. 4) where the sample groups corresponding to the sites were well separated from each other. it can also be seen that the first axis of the pcoa, as well as the first node of the classification tree, separated the mediterranean sites tar, esq and cha from the two other sites. the taxa preferring the epizoic habitat analysis of indicator values across sites revealed the preference of only two taxa for the epizoic habitat (indval=0.6; p≤0.06, tab. 2). these are navicula trivialis and fallacia pygmaea. the indval value of 0.6 resulted from the absence of both species in all epilithic or epiphytic samples and their presence in three out of five mixed epizoic samples. both species were present at the sites cha and esq. navicula trivialis was also found at sav and f. pygmaea at tar. intra-site differences between epizoic and non-epizoic communities the composition of the epizoic communities differed from that of the epilithic and epiphytic communities in table 2. indicator values (indval) of species significantly (p<0.1) associated with the epizoic habitat, at the scale of the five sites and in each site separately. site taxon indval p-value all sites navicula trivialis lange-bertalot 0.6 0.059 fallacia pygmaea (kützing) stickle & d.g.mann in round et al. 0.6 0.060 sav pseudostaurosira brevistriata (grunow) d.m.williams & round 0.97 0.018 staurosirella mutabilis (w.smith) e.morales & van de vijver in morales et al. 0.94 0.051 navicula cryptocephala kützing 0.90 0.069 staurosira venter (ehrenberg) cleve & j.d.möller 0.90 0.071 cyclotella distinguenda hustedt 0.82 0.045 staurosira binodis lange-bertalot in hofmann et al. 0.80 0.094 ver frustulia crassinervia (brébisson ex w.smith) lange-bertalot & krammer 0.86 0.086 sellaphora pupula (kützing) mereschkowksy 0.86 0.087 gomphonema auritum a.braun ex kützing 0.86 0.092 discostella stelligera var. tenuis (hustedt) houk & klee 0.77 0.021 aulacoseira pusilla (f.meister) tuji & houki 0.68 0.028 cha cocconeis placentula var. placentula ehrenberg 0.90 0.059 luticola goeppertiana (bleisch) d.g.mann ex rarick, s.wu, s.s.lee & edlund 0.88 0.012 esq bacillaria paxillifera (o.f.müller) t.marsson 1.00 0.016 navicula trivialis lange-bertalot 0.90 0.055 tryblionella hungarica (grunow) d.g.mann 0.97 0.027 tar navicula veneta kützing 0.73 0.028 fig. 4. plot of the principal coordinate analysis performed on individual epizoic samples (n=47). v. roubeix et al.6 only two of the five sites. a cut in the classification tree at distance 165 separated two groups of samples at each of the sites (fig. 3). these groups corresponded to substrate type (animal vs mineral/vegetal) at the sites ver and sav. at the other sites, substrate type did not seem to be an important factor. however, at the site cha, non-epizoic communities differed from a group of six epizoic samples but were closer to them than to a group of four other marginal epizoic samples. the analysis of indicator values at these sites revealed a significant ecological preference of some taxa for the animal substrate (tab. 2). at the site sav, these taxa were mostly of the staurosiraceae family. at the sites sav and ver, the epizoic habitat was characterized by planktonic species of the genera discostella, aulacoseira (ver) and cyclotella (sav). species of the bacillariaceae family obtained very high indicator values at the site esq (bacillaria paxillifera and tryblionella hungarica). at each site, some species of naviculales were also found indicator of turtle communities, such as navicula cryptocephala (sav), n. veneta (tar) or n. trivialis (esq), or frustulia crassinervia and sellaphora pupula (ver) or luticola goeppertiana (cha). a more detailed analysis of the communities was justified in the two sites where differences between substrates emerged (ver and sav). indeed, the variations in community composition between substrates observed at these two sites led to some discrepancies in the representation of ecological guilds and life forms. however, the patterns were not the same at the two sites. at the site sav, the functional compositions of diatom communities on stones and plants were similar (p>0.4) whereas the composition of epizoic diatoms differed significantly from them (p<10-15, fig. 5). there were more ‘high profile’ guild species in the epizoic communities compared to communities of other substrate types. at the site ver, the differences were significant between all substrates (p<10-6). there were more planktonic diatoms on the turtle carapaces (p≤10-6). considering life forms rather than ecological guilds, both sites had a lower proportion of pioneer diatoms on turtles, which was the only common feature of the epizoic communities at these sites, whereas the site sav was characterized by more colonial forms and the site ver by fewer pedunculated forms on turtles (fig. 6). discussion epizoic diatoms were present on the carapace of emys orbicularis, as previously observed on other freshwater (donato et al., 2018) and marine (robinson et al., 2016) turtle species. the numbers of species identified were higher than those reported in previous works (soylu et al., 2006; robinson et al., 2016; wu and bergey, 2017; donato et al., 2018). this can be explained by differences in sampling protocols or by the higher diversity of diatoms in the sites explored in this study. while small differences in specific and functional composition could be found between substrates, especially at two sites (ver and cha), it appeared that the carapace of the pond turtle did not harbour a consistent, characteristic diatom community. the differences in composition between the five sites were very significant and were probably due to local ecological factors and biogeographic effects, which in particular linked the three mediterranean sites (tar, esq and cha). this has to be compared with the study of american populations of another species of freshwater turtle, chelydra serpentina (wu and bergey, 2017). it was found that some epizoic communities of turtle populations, although very distant in space, were not significantly different. the diatoms present on marine turtles provide an even greater contrast to the present study. indeed, it appeared that marine turtles had on their carapace little diversified and similar diatom communities, whatever the species or the origin of the turtle individuals. moreover, some taxa were identified as obligatory epibionts (e.g., poulinea or chelonicola) (robinson et al., 2016). nevertheless, the homogeneity of epizoic diatom communities among turtle populations is not sufficient to demonstrate their specificity to the turtle carapace. it is necessary to examine in parallel the benthic communities present on other substrates in the vicinity of the turtle populations. at the scale of the five sites in south-eastern france and among the large number of taxa recorded on the fig. 5. proportions of the four diatom ecological guilds in the epilithic, epiphytic and mixed epizoic samples from the sites ver and sav (same codes as in figure 2). the letters above the plot indicate groups of significantly different compositions (p<0.05). diatoms on the european pond turtle carapace 7 pond turtle carapace, only two species were identified as characteristic of the epizoic habitat, while appearing in only three out of five sites (tab. 2). navicula trivialis and f. pygmaea are however commonly found on a variety of benthic substrates in the framework of aquatic ecosystem monitoring. they cannot be considered as typically epizoic species even though the turtle carapace appears to be a suitable substrate for them. they should be distinguished from the species luticola deniseae found exclusively on a brazilian freshwater turtle (wetzel et al., 2010) and from the marine genera poulinea and chelonicola considered as obligate epibionts (robinson et al., 2016). although no specific diatom communities or taxa could be found on the pond turtle carapace, different diatom assemblages on turtles, plants and stones were expected at the same site. indeed, compared to other substrates, the carapace presents particular characteristics that should select an adapted epizoic community. the mobility of the turtle generates variable light conditions and friction forces with water or objects that the animal comes into contact with. it should also reduce sediment build-up on the carapace. the movements of individuals to other water bodies (owen-jones et al., 2015; fuentes and olivier, 2016) may introduce new species not present in the sampling site. in addition, certain life history traits of the pond turtle should considerably influence epizoic communities. the turtles’ egg-laying out of the water promote exchanges with soil microorganisms. moreover, the ectotherm turtles get out of the water almost daily in spring to expose themselves to sunlight (basking) and this represents an important pressure on the biofilm of the carapace. it has even been hypothesized that this behaviour allows the turtles to control by desiccation the epizoic algal biomass that can harm their health when algal development becomes excessive (neil and allen, 1954). although it has not been demonstrated, turtles’ carapaces may contain chemicals that have antifouling properties (e.g., such as those found in marine sponges; amsler et al., 2000). however, turtle carapace communities differed in epilithic and epiphytic communities in only the two most inland sites (sav and ver) (fig. 3). the pattern could have been the same at the site cha if four marginal epizoic samples had not been taken into account. it was interesting to note the functional implications of the variations in composition between epizoic and non-epizoic substrates (figs. 5 and 6). the lower proportion of pioneer species suggests a more mature biofilm on turtle carapaces. although autochthonous mineral substrates were used for sampling at these sites, the age of the biofilm on pebbles or macrophytes may have been reduced by water level fluctuations or plant growth. the higher relative abundance of ‘high profile’ colonial and fig. 6. proportions of the main life forms of periphytic diatoms in the epilithic (lith), epiphytic (phy) and mixed epizoic (zmi) samples from the sites ver (a) and sav (b). (note: the same species may belong to two different groups). errors bars indicate 95% confidence intervals. v. roubeix et al.8 planktonic (unattached) species and the lower representation of pedunculated species suggest living conditions with less turbulence or abrasive forces (passy, 2007), perhaps distant from nearshore area where epilithic and epiphytic communities were collected. furthermore, the preponderance of planktonic forms indicates a deeper living environment where planktonic diatoms are more numerous and are deposited on benthic substrates. the observed differences in composition could therefore be explained by the frequentation by the turtles of deeper areas of the water body than those where stones and plants were found for sampling. other studies have shown quite different distributions of diatom life forms attached to turtles. in the study of seven freshwater turtle species, a common feature of epibiotic diatoms was the quasi absence of solitary species in favour of colonial forms (donato et al., 2018), while in the study of seven marine turtle species, adnate forms were dominant (robinson et al., 2016). although differences with other substrates were found for the sites sav and ver, there is apparently no universal functional community profile characterizing the turtle epizoic habitat. even if the compositions of the epizoic and non-epizoic communities were consistently not very different (case of the cha, tar and esq sites), the analysis of the indicator values could reveal certain taxa preferentially associated with turtle carapaces. it should be noted that no indicator taxon of epizoic habitat was common to two different sites (tab. 2). it therefore seems unlikely that an obligate epibiont could be found among the indicator species identified, especially since most of these species are common on other substrates or in the water column, in the case of planktonic diatoms (sites sav and ver). they are therefore facultative epibionts having found favourable conditions on the pond turtle carapace or surviving there until accessing other habitats. however, recent literature on diatoms attached to turtles draws attention to a particular taxon, listed as an indicator in the site cha: luticola goeppertiana (tab. 2). this species reached a relative abundance of 9 % in the mixed epizoic sample from the site. the genus luticola was the subject of particular interest in two studies dealing with diatom-turtle relationships in freshwater. a new species, l. deniseae, was discovered on the carapace of the turtle species podocnemis erythrocephala in brazil, with which it appeared to be specifically associated excluding other substrates (wetzel et al., 2010). in the usa, two luticola species were found dominant on all carapaces of the turtle chelydra serpentina, with a preponderance of luticola cf. goeppertiana (wu and bergey, 2017). it therefore seems that the genus luticola is particularly adapted to turtle carapaces, perhaps because of its resistance to desiccation (wu and bergey, 2017). the closest widespread species to l. deniseae is l. goeppertiana (wetzel et al., 2010). further taxonomic analysis of luticola species found on carapaces may indicate whether l. deniseae is an obligate epibiont of freshwater turtles that may occur on different host species across continents. conclusions overall, the epizoic diatom communities on the carapace of emys orbicularis identified at all five study sites are not specific, as they appear strongly influenced by the environment close to the turtles and are not or hardly distinguishable from the surrounding epilithic or epiphytic communities. this is consistent with the concept of facultative epibiosis (wahl and mark, 1999). at some sites, differences in specific and functional composition from mineral and plant substrates were observed, probably due to turtle-specific characteristics, such as those associated with their behaviour. however, these differences were site-specific and were not sufficient to demonstrate a general trend in the selection of taxa on turtle carapaces. at the species level, it was difficult to find taxa preferring epizoic habitat at the scale of the five sites. the carapace of the european pond turtle can be considered relatively neutral and could be used for sampling periphytic diatoms as well as stones or macrophytes for bioindication of water quality (vassal et al., 2020). this appears to be less the case for the american species chelydra serpentina (wu and bergey, 2017) and even less for marine turtle species (robinson et al., 2016). in the marine environment, the habitat of a turtle species, either neritic or pelagic, probably influences the specificity of the epizoic communities compared to the periphyton which covers the closest benthic substrates (frick and pfaller, 2013). similarly, in freshwater environments, the contact of a species with other substrates in the euphotic zone and the contact of individuals with each other might be important drivers. acknowledgments we are indebted to the many people who participated in the turtle catches, in particular maëlys marage (tour du valat), rozen rocher and cédric roy (cen paca), olivier scher (cen lr), joanny piolat (association lo parvi), florian véron and magalie rambourdin (cen allier). we thank the french ministry of environment for giving us the permission to capture european pond turtles at all sampling sites. this study was made possible thanks to the support of séverine lopez and the management of the service biodiversité eau et paysage of the dreal paca. we would like to express our sincere thanks to them. diatoms on the european pond turtle carapace 9 corresponding author: vincent.roubeix@ifremer.fr. present address: ifremer rbe/be, rue de l'ile d'yeu, bp 21105, 44311 nantes cedex 03, france keywords: emys orbicularis; epizoic habitat; microalgae; selection; specificity. received: 11 october 2020. accepted: 7 february 2021. this work is licensed under a creative commons attribution noncommercial 4.0 license (cc by-nc 4.0). ©copyright: the author(s), 2021 licensee pagepress, italy advances in oceanography and limnology, 2021; 12:9119 doi: 10.4081/aiol.2021.9119 references akgul r, caprazli t, turker n, erduğan h, tosunoğlu m, 2014. epizoic algae on emys orbicularis (linnaeus, 1758), and mauremys rivulata (valenciennes, 1833), in the kayak river delta (saros bay, turkey) (testudines: emydidae, geoemydidae). herpetozoa 27:21-28. amsler cd, moeller cb, mcclintock jb, iken kb, baker bj, 2000. chemical defenses against diatom fouling in antarctic marine sponges. biofouling 16:29-45. doi: 10.1080/ 08927010009378428 anderson mj, 2001. a new method for non-parametric multivariate analysis of variance. austral ecol. 26:32-46. doi: 10.1111/j.1442-9993.2001.01070.pp.x biggs bjf, stevenson rj, lowe rl, 1998. a habitat matrix conceptual model for stream periphyton. arch. hydrobiol. 143:21-56. doi: 10.1127/archiv-hydrobiol/143/1998/21 blindow i, 1987. the composition and density of epiphyton on several species of submerged macrophytes the neutral substrate hypothesis tested. aquat. bot. 29:157-168. doi: 10.1016/0304-3770(87)90093-3 cadi a, nemoz m, thienpont s, joly p, 2008. annual home range and movement in freshwater turtles: management of the endangered european pond turtle (emys orbicularis). rev. esp. herp. 22:71-86. di trani c, zuffi mal, 1997. thermoregulation of the european pond turtle emys orbicularis, in central italy. chelonian conserv. biol. 2:428-430. donato j, gonzález-trujillo j, romero b, castro rebolledo m, 2018. diatom assemblages associated with turtle carapaces in the neotropical region. rev. biol. trop. 66:1362-1372. doi: 10.15517/rbt.v66i4.31396 dufrêne m, legendre p, 1997. species assemblages and indicator species: the need for a flexible asymmetrical approach. ecol. monogr. 67:345-366. doi: 10.1890/00129615(1997)067[0345:saaist]2.0.co;2 edgren ra, edgren mk, tiffany lh, 1953. some north american turtles and their epizoophytic algae. ecology 34:733740. doi: 10.2307/1931336 ersanli et, gonulol a, 2014. epizoophyte composition on spotted turtle emys orbicularis (linnaeus, 1758) in turkey. bangladesh j. bot. 43:157-161. doi: 10.3329/bjb.v43i2. 21667 fayolle s, moriconi c, oursel b, koenig c, suet m, ficheux s, logez m, olivier a, 2016. epizoic algae distribution on the carapace and plastron of the european pond turtle (emys orbicularis, linnaeus, 1758): a study from the camargue, france. cryptogamie algol. 37:221-232. doi: 10.7872/crya/ v37.iss4.2016.221 frick m, pfaller j, 2013. sea turtle epibiosis, p. 399-426. in: j. wyneken, k.j. lohman, and j.a. musick (eds.), the biology of sea turtles. boca raton. fuentes t, olivier a, 2016. movement records of the european pond turtles (emys orbicularis) in the camargue (southern france). tour du valat. irstea, 2013. [protocole d’échantillonnage des communautés de phytobenthos en plans d’eau. mise en œuvre de la directive cadre européenne sur l’eau pour les plans d’eau].[in french]. irstea, bordeaux. available from: https://hydrobiodce.inrae.fr/telecharger/phytobenthos-plans-deau/ lange-bertalot h, hofmann g, werum m, cantonati m, 2017. freshwater benthic diatoms of central europe: over 800 common species used in ecological assessment. english edition with updated taxonomy and added species. koeltz botanical books, schmitten-oberreifenberg: 942 pp. majewska r, santoro m, bolaños f, chaves g, de stefano m, 2015. diatoms and other epibionts associated with olive ridley (lepidochelys olivacea) sea turtles from the pacific coast of costa rica. plos one 10:1-15. doi: 10.1371/journal.pone.0130351 majewska r, robert k, van de vijver b, nel r, 2020. a new species of lucanicum (cyclophorales, bacillariophyta) associated with loggerhead sea turtles from south africa. bot. lett. 167:7-14. doi: 10.1080/23818107.2019.1691648 naulleau g, 1992. study of terrestrial activity and aestivation in emys orbicularis (reptilia, chelonia) using telemetry, p. 343–346. in: z. korsos and i. kiss (eds.), proceedings of the 6th ordinary general meeting of the societas europaea herpetologica, budapest. neil wt, allen er, 1954. algae on turtles: some additional considerations. ecology 35:581-584. doi: 10.2307/1931051 oksanen j, blanchet fg, friendly m, kindt r, legendre p, mcglinn d, minchin pr, o’hara rb, simpson gl, solymos p, stevens mhh, szoecs e, wagner h, 2019. vegan: community ecology package. r package version 2.5-4. owen-jones z, priol p, thienpont s, cheylan m, sauret g, coïc c, besnard a, 2015. the contrasting effects of shortand long-term habitat drainage on the population dynamics of freshwater turtles in a human-dominated landscape. freshwater biol. 61:121-132. doi: 10.1111/fwb.12687 passy si, 2007. diatom ecological guilds display distinct and predictable behavior along nutrient and disturbance gradients in running waters. aquat. bot. 86:171-178. doi: 10.1016/ j.aquabot.2006.09.018 poulíčková a, manoylov k, 2019. ecology of freshwater diatoms – current trends and applications, p. 289–309. in: j. seckbach and r. gordon (eds.), diatoms: fundamentals and applications. john wiley & sons, ltd. doi: 10.1002/ 9781119370741.ch13 rhodin ag, stanford cb, van dijk pp, eisemberg c, luiselli l, mittermeier ra, et al, 2018. global conservation status v. roubeix et al.10 of turtles and tortoises (order testudines). chelonian conserv. biol. 17:135-161. doi: 10.2744/ccb-1348.1 rimet f, bouchez a, 2012. life-forms, cell-sizes and ecological guilds of diatoms in european rivers. knowl. manag. aquat. ec. 406:1. doi: 10.1051/kmae/2012018 roberts dw, 2019. labdsv: ordination and multivariate analysis for ecology. robinson nj, majewska r, lazo-wasem ea, nel r, paladino fv, rojas l, zardus jd, pinou t, 2016. epibiotic diatoms are universally present on all sea turtle species. plos one 11:1-8. doi: 10.1371/journal.pone.0157011 rovero f, chelazzi g, 1996. nesting migrations in a population of emys orbicularis (l.) (chelonia emydidae) from central italy. ethol. ecol. evol. 8:297-304. doi: 10.1080/ 08927014.1996.9522920 soylu e, gönülol a, sukatar a, ayaz d, tok c, 2006. epizoic freshwater algae on emys orbicularis (testudinata: emydidae) from the central anatolia region of turkey. j. freshwater ecol. 21:535-538. doi: 10.1080/02705060.006. 9665033 thienpont s, 2020. plan national d’actions en faveur de la cistude d’europe (emys orbicularis) 2020-2029. société herpétologique de france. ministère de la transition ecologique. 103 pp. vassal v, ector l, van de vijver b, roubeix v, olivier a, pauvert s, roy c, fayolle s, 2020. pond turtle carapaces, an alternative natural substrate for the use of a diatom-based water quality index. bot. lett. doi: 10.1080/23818107.2020.1724825 vidakovic d, dimitrijevic j, krizmanic i, krizmanic j, 2018. diatoms on the carapace of european pond turtle (emys orbicularis) in vojvodina. proceedings of the vii balkan botanical congress, novi sad. wahl m, 1989. marine epibiosis. i. fouling and antifouling: some basic aspects. mar. ecol. prog. ser. 58:175-189. doi: 10.3354/meps058175 wahl m, 2008. ecological lever and interface ecology: epibiosis modulates the interactions between host and environment. biofouling 24:427-438. doi: 10.1080/08927010802339772 wahl m, mark o, 1999. the predominantly facultative nature of epibiosis: experimental and observational evidence. mar. ecol. prog. ser. 187:59-66. doi: 10.3354/meps187059 wetzel ce, van de vijver b, ector l, 2010. luticola deniseae sp. nov. a new epizoic diatom from the rio negro (amazon hydrographic basin). vie milieu 60:177-184. wetzel ce, van de vijver b, cox ej, bicudo dc, ector l, 2012. tursiocola podocnemicola sp. nov., a new epizoic freshwater diatom species from the rio negro in the brazilian amazon basin. diatom res. 27:1-8. doi: 10.1080/0269249x. 2011.642498 wu sc, bergey ea, 2017. diatoms on the carapace of common snapping turtles: luticola spp. dominate despite spatial variation in assemblages. plos one 12:e0171910. doi: 10.1371/ journal.pone.0171910 layout 1 introduction subfossil cladocera (crustacea, branchiopoda) represent one of the most valuable biological proxies preserved in lake sediments that can be studied for reconstruction purposes (kohrola and rautio, 2011). they are widespread in both the pelagic and littoral zones of lakes of different geographical distribution, altitude and typology, where they often represent the dominant component of zooplankton in terms of biomass. the chitinous parts of their body are well preserved in lake sediments, and the taphonomic taxonomy is well established, thanks to the numerous studies that followed the first pioneer works by frey (1960). cladocera play a key ecological role in freshwater ecosystems, as they occupy an intermediate position in the food web between primary producers (phytoplankton) and primary consumers (invertebrates and fish). as a consequence, subfossil cladocera remains have the capability to track long term changes in both bottom-up drivers (such as nutrients, physical and chemical stressors) and top down regulators, such as invertebrate and fish predation (e.g., jeppesen et al., 2001; szeroczyńska, 2006; perga et al., 2015). the changes in taxonomical composition of subfossil cladocera, which mainly includes bosminidae and chydoridae, and secondly daphniidae, have been increasingly investigated during the last decades and successfully used to track past environmental changes related to nutrient enrichment (lotter et al., 1998; bigler et al. 2007; manca et al., 2007; nevalainen and luoto, in press), acidification and calcium decline (krause and dellin, 1986; paterson, 1994; jeziorski et al., 2008), chemical contamination (korosi and smol, 2012a; labaj et al., 2016), hydrological changes (korhola et al., 2005; nevalainen et al., 2011), submerged macrophytes (davidson et al., 2011a), and climate change (lotter et al., 1997; kamenik et al., 2007; korponai et al., 2011; nevalainen et al., 2013; zawiska et al., 2015). the strong response of cladocera remains to environmental variability led to inference methods for quantitative reconstruction of past lake water variables, especially phosphorus (brodersen et al., 1998), lake depth (davidson et al., 2011b; nevalainen et al., 2011), and water temperature (duigan and birks 2000; lotter et al., 2000). in addition, subfossil cladoceran remains preserved in lakes sediments have the very valuable capability to allow reconstructing past changes in the lake food-web induced by the predation pressure by planktivorous fish (e.g., korosi et al., 2013). information on past fish populations and predation pressure on lacustrine zooplankton is in many case scattered, partial, or controversial, as it often relies on imprecise historical data, or on catch records from sport or commercial fisheries, the latter being biased by temporal changes in the catches of certain species due to their fluctuating commercial value. within this context, changes in species composition and abundances of cladocera remains can support the indirect reconstruction of food web changes in both temperate and high altitude/high latitude lakes (which are mainly naturally fishless, but experienced historical legal or illegal fish introductions), thus fostering conservation and restoration actions (e.g., tiberti et al., 2014). recent investigations revealed that not only species composition and abundance, but also morphology of cladoceran remains can be used for ecological reconstructions. it is well established that invertebrate and fish predation can affect body size, morphology and pigmentation of cladocera (jeppesen et al., 2002; hansson, 2004; guilizzoni et al., 2006;). however, pigmentation has been recently used also to track changes in underwater uv radiation in relation to solar activity (nevalainen and rautio, 2014) and changes in water doc concentrations, the latter in relation to lake productivity (nevalainen et al, 2016) or changing land use within the lake catchment (e.g., advances in oceanography and limnology, 2016; 7(2): 125-130 doi: 10.4081/aiol.2016.6467 this work is licensed under a creative commons attribution-noncommercial 4.0 international license (cc by-nc 4.0). subfossil cladocera as a powerful tool for paleoecological reconstruction monica tolotti,1* manuela milan,2 krystyna szeroczyńska3 1sustainable agro-ecosystems and bioresources, research and innovation centre (cri), fondazione e. mach, via e. mach 1, 38010 san michele all'adige, tn, italy 2department of ecology and environmental sciences (emg), umeå university, linnaeus väg 6, 90187 umeå, sweden 3institute of geological sciences, polish academy of sciences, research centre warsaw, twarda 51/55, pl 00818 warsaw, poland *corresponding author: monica.tolotti@fmach.it key words: cladocera, paleoecology, lakes, sediments, human impact, climate change. received: december 2016. accepted: december 2016. no n c om me rci al us e o nly 126 m. tolotti, m. milan, and k. szeroczyńska a-forestation, water regulation). isotopic composition of cladoceran remains also revealed to be a very time-effective and promising tool for interpreting changes in lake food web and functionality (perga et al., 2010; perga, 2011). nevertheless, a set of factors still hamper the interpretation of sedimentary cladocera results, and, in turn, the exploitation of the great potential of this biological proxy for the reconstruction of past lake evolution. firstly, taxonomy of subfossil cladocera is well established for temperate and boreal regions of europe and north america (szeroczyńska and sarmaja-korionen, 2007; korosi and smol, 2012b), whereas only a few works has been published for tropical regions (cuna et al., 2014; sinev and zawisza, 2013; szeroczyńska et al., 2015). much work remains to be done also to evaluate how well taphonomic cladocera represent the living communities (kattel et al., 2007; alric and perga, 2011; kirillova et al., 2016), and how preservation of remain type and species in the sediment can be affected by water characteristics, such as oxygenation, ph and chemical composition. sedimentation dynamics can also affect spatial distribution of remains in the lake sediments (alric and perga, 2011). finally, the interpretation of sediment records is complicated by the reciprocal interactions between multiple drivers, such as climate and nutrients, which can produce additive, competitive or synergic effects (battarbee et al., 2012). although the multi-proxy paleoecological approach and the comparison of sediment records with limnological data (e.g. manca et al., 2007; bennion et al., 2015) can help disentangling the effects of multiple drivers, more studies are still necessary to make the interpretation of the cladocera sediment records more straightforward. the first subfossil cladocera workshop was initiated by prof. atte korhola in 1999 (helsinki) with the aim of getting together specialists, young researchers and students working on various aspects of cladocera remains in lake sediments in order to share knowledge, to foster discussion on new ecological findings and ideas, and to practice species identification at the microscope under the guidance of expert taxonomists. the xiv subfossil cladocera workshop was organized within this same spirit and held at levico terme (italy) from 5th to 8th april 2016. the 30 participants (fig. 1) from 9 countries (czech republic, finland, france, germany, hungary, italy, poland, russia, uk) were almost equally distributed between senior, young scientists and students, what promoted the transfer of knowledge and experience among generations. one special objective of this workshop was to stimulate the discussion on possible future developments of cladocera-based paleoecological reconstructions based on relatively new approaches, such as the “resurrection ecology” techniques, the study of isotopic signatures in body and ephippia remains, and the statistical treatment of multiproxy sediment data. thematic papers the thematic papers grouped in this volume represent the outcome of presentations and discussions held at the xiv subfossil cladocera workshop. the contributions focus on taxonomy, diversity, distribution of cladocera remains in lake sediments in europe and america, as well as on the subfossil cladocera capability to track past changes in both bottom-up and top-down drivers of lake ecological dynamics. several papers aimed at reconstructing the environmental evolution of temperate european lakes at secular (milan et al., 2016) or millennial (niska, 2016; szeroczyńska, 2016; zawisza et al., 2016) scales. the contribution by milan et al., (2016) showed how the multiproxy approach, and in particular the combination of biological proxies and geochemical analyses, could improve the understanding of the relation between cladoceran communities and hydrological variability. the paleolimnological studies at millennial scale investigated the relations between cladocera and environmental drivers in stages where human impact was still absent or negligible, thus allowing the discrimination of climate related effects. korponai et al. (2016) used subfossil cladocera to distinguish lentic and lotic stages in oxbow lakes along the river tisza (hungary), thus demonstrating the potential of cladoceran remains to reconstruct changes in the hydrological regimen of transitional water ecosystems. the thematic section of this volume tackled also the taxonomic issue. wojewódka et al. (2016) presented a first description of cladocera diversity in superficial sediments of 29 lakes of different altitude and size in central america, thus contributing to the improvement of the cladoceran taxonomy within this still scarcely investigated region. on the other side, zawiska et al. (2016) described a time and cost effective method to prepare subfossil cladocera for sem analysis, which allows the observation of taxonomically important details of the structure and ornamentation of carapace and spines. finally, several contributions studied the importance of morphological variability of cladocera remains in tracking long term environmental and ecological changes. leppänen and weckström (2016) explored the potential use of changes in size and preservation level of daphnia caudal spines to track fishing and forestry activities, as well as changes in water ph. milan et al. (2016) analyzed changes in bosminidae and daphniidae body size and appendages length to reconstruct major changes in the lake food-web, outlining nutrient enrichment and appearance of predator cladocera species as the major drivers of size changes. szeroczyńska (2016) related the presence of extreme eubosmina morphs, observed in a german lake, to stages of particularly pronounced water turbulence and turbidity. finally, bérubé tellier et al. (2016) demonstrated as variano n c om me rci al us e o nly subfossil cladocera for paleoecology 127 tions in resting eggs (ephippia) pigmentation can be used to track changes in predation pressure by fish and changes in water doc concentration in relation to catchment land use. in conclusion, the thematic papers included in this volume consolidate the indicator value of subfossil cladocera in both classical and more novel paleolimnological approaches, outlining at the same time several research topics, which need to be further developed in the near future. fig. 1. the participants of the xiv subfossil cladocera workshop during the excursion at lake garda. no n c om me rci al us e o nly 128 m. tolotti, m. milan, and k. szeroczyńska references alric b, perga me, 2011. effects of production, sedimentation and taphonomic processes on the composition and size structure of sedimenting cladoceran remains in a large deep subalpine lake: paleo-ecological implications. hydrobiologia 676:101-116. battarbee rw, anderson nj, bennion h, simpson gl, 2012. combining limnological and palaeolimnological data to disentangle the effects of nutrient pollution and climate change on lake ecosystems: problems and potential. freshwater biol. 57:2091-2106. bennion h, davidson ta, sayer cd, simpson g, rose nl, sadler j, 2015. harnessing the potential of the multi-indicator palaeolimnological approach: an assessment of the nature and causes of ecological changes in a eutrophic shallow lake. freshwater biol. 60:1423-1442. bérubé tellier a, drevnick pe, bertolo a, 2016. brook trout (salvelinus fontinalis) extinction in small boreal lakes revealed by ephippia pigmentation: a preliminary analysis. adv. oceanol. limnol. 7:6215. bigler c, von gunten l, lotter af, hausmann s, blass a, ohlendorf c, sturm m, 2007. quantifying human-induced eutrophication in swiss mountain lakes since ad 1800 using diatoms. holocene 17:1141-1154. brodersen kp, whiteside mc, lindegaard c, 1998. reconstruction of trophic state in danish lakes using subfossil chydorid (cladocera) assemblages. can. j. fish. aquat. sci. 55:10931103. cuna e, zawisza e, caballero m, ruiz-fernández ac, lozanogarcía s, alcocer j, 2014. environmental impacts of little ice age cooling in central mexico recorded in the sediments of a tropical alpine lake. j. paleolimnol. 51:1-14. davidson ta, sayer c, perrow m, bramm m, jeppesen e, 2007. are the controls of species composition similar for contemporary and sub-fossil cladoceran assemblages? a study of 39 shallow lakes of contrasting trophic status. j. paleolimnol. 38:117-134. davidson ta, bennion h, jeppesen e, clarke gh, sayer cd, morley d, odgaard bv, rasmussen p, rawcliffe r, salgado j, simpson gl, amsinck sl, 2011a. the role of cladocerans in tracking long-term change in shallow lake trophic status. hydrobiologia 676:299-315. davidson t a, amsinck sl, bennike o, christoffersen k, landkildehus f, lauridsen tl, jeppesen e, 2011b. inferring a single variable from assemblages with multiple controls: getting into deep water with cladoceran lake depth transfer functions. hydrobiologia 676:299-315. duigan ca, birks hh, 2000. the late-glacial and earlyholocene palaeoecology of cladoceran microfossil assemblages at kråkenes, western norway, with a quantita tive reconstruction of temperature changes. j. paleolimnol. 23:67-76. frey d g, 1960. the ecological significance of cladoceran remains in lake sediments. ecology 41:684-699. frossard v, verneaux v, milelt l, jenny jp, arnaud f, magny m, perga me, 2014. reconstructing long-term changes (150 years) in the carbon cycle of a clear-water lake based on the stable carbon isotope composition (d13c) of chironomid and cladoceran subfossil remains. freshwater biol. 59:789-802. guilizzoni p, lami a, manca m, musazzi s, marchetto a, 2006. palaeoenvironmental changes inferred from biological remains in short lake sediment cores from the central alps and dolomites. in: a. lami and a. boggero (eds.), ecology of high altitude aquatic systems in the alps. hydrobiologia 562:167-191. hansson la, 2004. plasticity in pigmentation induced by conflicting threats from predation and uv radiation. ecology 85:1005-1016. jeppesen e, leavitt p, de meester l, jensen jp, 2001. functional ecology and palaeolimnology: using cladoceran remains to reconstruct anthropogenic impact. trends ecol. evol. 16:191-198. jeppesen e, jensen jp, amsinck s, landkildehus f, lauridsen t, mitchell sf, 2002. reconstructing the historical changes in daphnia mean size and planktivorous fish abundance in lakes from the size of daphnia ephippia in the sediment. j. paleolimnol. 27:133-143. jeppesen e, nõges p, davidson ta, haberman j, nõges t, blank k, lauridsen tl, søndergaard m, sayer c, laugaste r, johansson ls, bjerring r, amsinck sl, 2011. zooplankton as indicators in lakes: a scientific-based plea for including zooplankton in the ecological quality assessment of lakes according to the european water framework directive (wfd). hydrobiologia 676:279-297. jeziorski a, yan nd, paterson am, desellas am, turner ma, jeffries ds, keller b, weeber rc, mcnicol dk, palmer me, 2008. the widespread threat of calcium decline in fresh waters. science 322:1374-1377. kamenik c, szeroczyńska k, schmidt r, 2007. relationships among recent alpine cladocera remains and their environment: implications for climate-change studies. hydrobiologia 594:33-46. kattel gr, battarbee rw, mackay a, birks hb, 2007. are cladoceran fossils in lake sediment samples a biased reflection of the communities from which they are derived? j. paleolimnol. 38:157-181. kattel gr, battarbee rw, mackay a, birks hjb, 2008. recent ecological change in a remote scottish mountain loch: an evaluation of a cladocera-based temperature transfer-function. palaeogeogr. palaeoclim. palaeoecol. 259:51-76. kirillova iv, van der plicht j, gubin sv, zanina og, chernova of, lapteva eg, trofimova ss, zinovyev ev, zharov aa, fadeeva eo, van kolfschoten t, shidlovskiy fk, kotov aa, 2016. taphonomic phenomenon of ancient hair from glacial beringia: perspectives for palaeoecological reconstructions. boreas 45:455-469. korhola a, rautio m, 2001. cladocera and other branchiopod crustaceans, p. 5-41. in: j.p. smol, h.j.b. birks and w.m. last (eds.), tracking environmental change using lake sediments, zoological indicators. 4. kluwer, dordrecht. korhola a, tikkanen m, weckström j, 2005. quantification of holocene lake-level changes in finnish lapland using a cladocera–lake depth transfer model. j. paleolimnol. 34:175-190. korosi jb, smol jp, 2012a. examining the effects of climate change, acidic deposition, and copper sulphate poisoning on long-term changes in cladoceran assemblages. aquat. sci. 74:781-792. korosi jb, smol jp, 2012b. an illustrated guide to the identifino n c om me rci al us e o nly subfossil cladocera for paleoecology 129 cation of cladoceran subfossils from lake sediments in northeastern north america: part 1 the daphniidae, leptodoridae, bosminidae, polyphemidae, holopedidae, sididae, and macrothricidae. j. paleolimnol. 48:571-586. korosi jb, kurek j, smol, jp, 2013. a review on utilizing bosmina size structure archived in lake sediments to infer historic shifts in predation regimes. j. plankton res. 35:444460. korponai j, magyari em, buczkó k, iepure s, namiotko t, czakó d, kövér c, braun m, 2011. cladocera response to late glacial to early holocene climate change in a south carpathian mountain lake. hydrobiologia 676:223-235. korponai j, gyulai i, braun m, kover c. papp i, forro l, 2016. reconstruction of flood events in an oxbow lake (marótzugi-holt-tisza, ne ungary) by using subfossil cladoceran remains and sediments. adv. oceanol. limnol. 7:6168. krause-dellin d, steinberg c, 1986. cladoceran remains as indicators of lake acidification. hydrobiologia 143:129-134. labaj al, korosi jb, kurek j, jeziorski a, keler wb, smol jp, 2016. response of bosmina size structure to the acidification and recovery of lakes near sudbury, canada. j. limnol. 75:22-29. leppänen jj, weckström, j, 2016. varying degradation of subfossil daphnia longispina during the past 250 years and the discovery of fossil helmet-type head shields: preliminary results. adv. oceanol. limnol. 7:6293. lotter af, birks hjb, hofmann w, marchetto a, 1997. modern diatom, cladocera, chironomid, and chrysophyte cyst assemblages as quantitative indicators for the reconstruction of past environmental conditions in the alps. i. climate. j. paleolimnol. 18:395-420. lotter af, birks hjb, hofmann w, marchetto a, 1998. modern diatom, cladocera, chironomid, and chrysophyte cyst assemblages as quantitative indicators for the reconstruction of past environmental conditions in the alps. ii. nutrients. j. paleolimnol. 19:443-463. lotter a, birks h, eicher u, hofmann w, schwander j, wick l, 2000. younger dryas and allerød summer temperatures at gerzensee (switzerland) inferred from fossil pollen and cladoceran assemblages. palaeogeogr. palaeoclim. palaeoecol. 159:349-361. lynn dv, hann bj, paterson m, 2015. littoral cladoceran community reassembly following the cessation of disturbance. j. paleolimnol. 54:121-135. manca m, torretta b, comoli p, amsinck sl, jeppesen e, 2007. major changes in trophic dynamics in large, deep sub-alpine lake maggiore from 1940s to 2002: a high resolution comparative palaeo-neolimnological study. freshwater biol. 52:2256-2269. milan m, bindler r, tolotti m, 2016. combining sediment cladocera remains and geochemistry to reveal the role of a large catchment in driving changes in a small subalpine lake (lake ledro, n-italy). adv. oceanol. limnol. 7:6399. nevalainen l, sarmaja-korjonen k, luoto tp, 2011. sedimentary cladocera as indicators of past water-level changes in shallow northern lakes. quaternary res. 75:430-437. nevalainen l, luoto tp, kultti s, sarmaja-korjonen k, 2013. spatio-temporal distribution of sedimentary cladocera (crustacea: branchiopoda) in relation to climate. j. biogeogr. 40:1548-1559. nevalainen l, rautio m, 2014. spectral absorbance of benthic cladoceran carapaces as a new method for inferring past uv exposure of aquatic biota. quaternary sci. rev. 84:109-115. nevalainen l, luoto tp (2016). relationship between cladoceran (crustacea) functional diversity and lake trophic gradients. funct. ecol. doi:10.1111/1365-2435.12737 (in press). nevalainen l, rantala mv, luoto tp, ojala ae, rautio m, 2016. long-term changes in pigmentation of arctic daphnia provide potential for reconstructing aquatic uv exposure. quaternary sc. rev. 144:44-50. niska m, 2016. the eemian/early vistulian development of the solniki paleolake (north-eastern poland) as shown by subfossil cladocera. adv. oceanol. limnol. 7:6217. paterson mj, 1994. paleolimnological reconstruction of recent changes in assemblages of cladocera from acidified lakes in the adirondack mountains (new york). j. paleolimnol. 11:189-200. perga me, 2011. taphonomic and early diagenetic effects on the c and n stable isotope composition of cladoceran remains: implications for paleoecological studies. j. paleolimnol. 46:203-213. perga me, desmet m, enters d, reyss jl, 2010. a century of bottom-upand top-down-driven changes on a lake planktonic food web: a paleoecological and paleoisotopic study of lake annecy, france. limnol. oceanogr. 55:803-816. perga me, frossard v, jenny jp, alric b, arnaud f, berthon v, black jl, domaizon i, giguet-covex c, kirkham a, magny m, manca m, marchetto a, millet l, paillès c pignol c, poulenard j, reyss jl, rimet f, sabatim p, 2015. high-resolution paleolimnology opens new management perspectives for lakes adaptation to climate warming. front. ecol. evol. 3:1-17. sinev a, zawisza e, 2013. comments on cladocerans of crater lakes of the nevado de toluca volcano (central mexico), with the description of a new species, alona manueli sp. nov. zootaxa 3647:390-400. szeroczyńska k, 2006. the significance of subfossil cladocera in stratigraphy of late glacial and holocene. studia quaternaria 23:37-45. szeroczyńska k, sarmaja-korjonen k, 2007. atlas of subfossil cladocera from central and northern europe. friends of the lower vistula society, świecie: 84 pp. szeroczyńska k, zawisza e, wojewódka m, 2015. initial time of two high altitude crater lakes (nevado de toluca, central mexico) recorded in subfossil cladocera. studia quaternaria 32:109-116. szeroczyńska k, 2016. long term subfossil cladocera record from the partly varved sediment of lake tiefer (ne germany). adv. oceanol. limnol. 7:6297. tiberti r, von hardenberg a, bogliani g, 2014. ecological impact of introduced fish in high altitude lakes: a case of study from the european alps. hydrobiologia 724:1-19. wojewódka m, zawisza e, cohuo s, macario-gonzález l, schwalb a, zawiska i, perez l, 2016. ecology of cladocera species from central america based on subfossil assemblages. adv. oceanol. limnol. 7:6266. zawiska i, słowiński m, correa-metrio a, obremska m, luoto tp, nevalainen l, woszczyk m, milecka k, 2015. the reno n c om me rci al us e o nly 130 m. tolotti, m. milan, and k. szeroczyńska sponse of a shallow lake and its catchment to late glacial climate changes a case study from eastern poland. catena 126:1-10. zawiska i, zawisza e, wojewódka m, sinev ay, 2016. exploring the world of micro sculptures subfossil cladocera remains under the sem. adv. oceanol. limnol. 7:6218. zawisza e, filbrandt-czaja a, correa-metrio a, 2016. subfossil cladocera and pollen as indicators of natural and anthropogenic trophic changes of lake jelonek (tuchola forest, n poland) during the holocene. adv. oceanol. limnol. 7:157-170. no n c om me rci al us e o nly layout 1 introduction marine diatoms are some of the most successful groups of unicellular eukaryotic photosynthetic organisms, forming massive blooms in coastal and open waters and having a key role in global marine primary production (armbrust, 2009). elucidating the mechanisms that regulate diatom bloom dynamics and mainly their decay determinants is therefore fundamental to understand nutrient and carbon cycling so as energy flow at the global as well as the local scale. programmed cell death (pcd) is an active, self-controlled death mechanism well known in higher organisms which has been also detected in unicellulars (bidle et al., 2007; segovia, 2011). however, its occurrence in unicellular organisms is still considered a matter of ongoing debate (deponte, 2008; jimenez et al., 2009), and the specific mode of action in which pcd is initiated remains to be elucidated. pcd has been suggested to act in phytoplankton populations to eliminate aged cells and to reduce population size in order to cope with nutrient limitation, for instance, in the final stages of blooms (bidle and falkowski, 2004; vardi et al., 1999, 2006, 2007). after a certain threshold of limitation of environmental factors (mainly nutrients), pcd induces the massive decay of the bloom, sometimes through massive production of resting or resistance stages (ellegaard et al., 2013; mcquoid et al., 2002). these factors include nitrogen and light limitation (berges and falkowski, 1998), iron starvation (bidle and bender, 2008; thamatrakoln et al., 2012), virus infection (bidle et al., 2007; vardi et al., 2009), co2 limitation (vardi et al., 1999) and several secondary metabolites (cervia et al., 2009; costas et al., 1993). reactive oxygen species (ros) are often involved in inducing pcd in reaction to a stress factor in a wide range of organisms and cell lines, from bacteria to mammalians and plants (andrianasolo et al., 2007; jones, 2008; kuwabara et al., 2008; lam, 2008; rockenfeller and madeo, 2008; scherz-shouval and elazar, 2007), mainly at intermediate concentrations. in fact, several cell reactions to ros strongly depend upon the ros concentration and may range from production of antioxidant compounds to necrotic cell death (chandra et al., 2000), or pcd (costas et al., 1993; vardi et al., 2009). apart from ros concentrations, several variables come into play by modulating and regulating the cell response to ros. these include the cell type, the stress type, its magnitude and its duration in time (martindale and holbrook, 2002). different case studies report the co-occurrence of ros and pcd in marine phytoplankton and suggest a correlation between them (butow et al., 1997; segovia and berges, 2009; vardi et al., 1999). in the dinoflagellate peridinium gatunense, pcd leading to the collapse advances in oceanography and limnology, 2015; 6(1/2): 13-20 original article doi: 10.4081/aiol.2015.5466 expression of death-related genes and reactive oxygen species production in skeletonema tropicum upon exposure to the polyunsaturated aldehyde octadienal alessandra a. gallina,1# chih-ching chung,2,3 raffaella casotti1* 1stazione zoologica anton dohrn di napoli, villa comunale, 80121 naples, italy; 2institute of marine environmental chemistry and ecology, national taiwan ocean university, keelung 20224, taiwan, china; 3center of excellence for the oceans, national taiwan ocean university, keelung 20224,taiwan, republic of china #present address: department of biology, colorado state university, fort collins, co 80523, usa *corresponding author: raffaella.casotti@szn.it abstract the effects of 4e/z-octadienal (octa) on scdsp-1 and scdsp-2 gene expression and reactive oxygen species (ros) production were investigated in the marine diatom skeletonema tropicum (formerly costatum) using qrtpcr and flow cytometry. scdsp-1 and scdsp-2 genes have been previously shown to be involved in cell death in ageing cells and in response to photosynthetic stress. octa induced a differential, concentration-dependent dsp gene expression associated to ros production, 821.6 and 97.7 folds higher for scdsp1 and scdsp-2, respectively. among the concentrations tested, only 8 μm octa, which caused a reduction of 50% in cell concentrations at 24 h, was able to elicit an expression pattern consistent with a signalling role. interestingly, only intermediate levels of reactive oxygen species (ros) (i.e., 1.5±0.1 increase) were observed to be elicited by such concentration. these results suggest that ros are key components of the molecular cascade triggered by polyunsaturated aldehydes (pua) and leading to cell death. this could have implications for bloom final stages at sea, where pua may act as effectors of diatom population dynamics through ros acting as modulators. key words: diatom; chemical ecology; bloom evolution; programmed cell death; ros. received: july 2015. accepted: october 2015. no nco mm er cia l u se on ly 14dsp expression in response to pua of the algal population is mediated by oxidative stress induced by carbon dioxide (co2) limitation (vardi et al., 1999). similarly, ros-mediated cell death was observed at the end of a bloom of the same species, despite an increase in antioxidant activity (butow et al., 1997). pcdlike cell death concomitant with accumulation of ros is also reported to be induced by the toxic cyanobacterium microcystis sp. (vardi et al., 2002), or h2o2 (vardi et al., 2007). in the chlorophyte dunaliella tertiolecta, dark-induced pcd, involving the activity of caspase-like enzymes has been linked to an increase in ros production (segovia and berges, 2009). in the prymnesiophyte emiliana huxley viral infection has been associated to an enhanced production of ros, possibly leading to pcd (evans et al., 2006). interestingly, the secondary metabolite euplotin c derived from the marine ciliated protist euplotes crassus, is found to induce pcd in the congeneric euplotes vannus, which does not produce this metabolite (cervia et al., 2009). this suggests that specific secondary metabolites can play an ecological role in broadening phytoplankton niche size through different mechanisms, including pcd (cervia et al., 2009). in the marine diatom phaeodactylum tricornutum, the exposure to diatom-derived secondary metabolites, polyunsaturated aldehydes (pua), results in an altered expression of metacaspases through the overexpression of a gene associated to no generation (vardi et al., 2008). in the marine diatom thalassiosira weissflogii, pua were shown to induce pcd-like cell death, suggesting for a role in activating the cell death cascade, therefore regulating population dynamics and species succession (casotti et al., 2005). in phaeodactylum tricornutum it is speculated that a sophisticated stress surveillance system exists, involving no and calcium, in response to pua, which could lead to bloom termination and population-level cell death at specific pua concentrations (vardi et al., 2006). even pua precursors, the polyunsaturated fatty acids (pufas), have been reported to inhibit growth of candida albicans and candida dubliniensi by inducing ros accumulation and consequent pcd (thibane et al., 2012). in the marine diatom skeletonema tropicum (formerly costatum) two genes have been identified, coding for two identical ca-regulated proteins, named s. costatum deathspecific proteins (scdsp-1 and scdsp-2) (chung et al., 2005), which are both composed of a transmembrane domain and a pair of ef-hand motifs (luan et al., 2002; means and dedman, 1980). both scdsp gene expressions were shown to be strongly upregulated in aging cells and also during light stress. their expression is correlated to dna fragmentation, suggesting for a possible role of the encoded protein in the signal transduction of stress to the cell death machinery (chung et al., 2005, 2008). in particular, scdsp expression has been related to the blockage of the electron flow between psii and cytochrome b6f and it was found to be no dependent (chung et al., 2008). moreover, two dsp-like proteins (tpdsp1 and tpdsp2), have been recently identified in the marine diatom thalassiosira pseudonana and have been shown to be upregulated both under iron limitation and starvation, suggesting for a role in fe-induced stress and consequent death (thamatrakoln et al., 2012). however, they have also recently been shown to enhance growth under fe limitation and to be controlled by light intensity (thamatrakoln et al., 2013). in this study we report the gene expression of the two death-related genes scdsp-1 and scdsp-2 in st exposed to the polyunsaturated aldehyde octadienal, and relate it to ros production. the overall aim is to relate cell death to chemical stress induced by pua and to advance hypotheses on bloom decay mechanisms at sea. methods experimental setup and cultures a unialgal culture of skeletonema tropicum cleve (formerly costatum) was grown at 23°c on a 12h-12h lightdark cycle under a photon flux density of 110 µmol quanta m–2 s–1 (chung et al., 2005). natural seawater, amended with f/2 nutrients (guillard, 1975) was used as medium. for gene expression measurements, s. tropicum was grown in 10 l polycarbonate carboys (nalgene). when the culture reached the late exponential phase of growth (le, fig. 1), cells were transferred into 2 l polycarbonate bottles (nalgene) and treated with different concentrations of 4e/z-octadienal (octa) or h2o2. samples were collected at 1,2,3,4,5 and 1,2,3,4,12 days since the beginning of the fig. 1. effect of octa and h2o2 on cell concentrations of s. tropicum. le is the time point indicated as late exponential growth phase. data are means of triplicates ±sd. when the bar is not visible, it is smaller than the symbol. no nco mm er cia l u se on ly 15 a.a. gallina et al. experiment for cell counts and total rna extraction, respectively. cell concentrations were determined by using a sedgwick-rafter counting chamber (hausser scientific partnership). at least 200 cells from each sample were counted using a light microscope (bx60; olympus, center valley, usa) at 200× magnification. growth rates were calculated using the following equation: µ=ln [(n1/n0)/t] (eq. 1) where n0 and n1 are cell density at the start and the end of the growth period, and t is the time between measurements (in days). since no dead vs alive cells assay has been used, it is possible that the cells counts reported in the study might include damaged or dying cells. gene expression experiments were performed using three octa concentrations, namely the concentration which caused a reduction of 50% in cell concentration at 24 h, half this concentration and twice this concentration (4, 8 and 16 µm). for ros detection, three independent cultures were grown in 1 l glass flasks until the late exponential growth phase. cells were harvested and directly exposed to the same concentrations of octa used for gene expression measurements and analysed through flow cytometry as described below. the controls were dyeloaded samples processed exactly as the other samples except for pua inoculation. each experiment was performed at least twice, but generally three times, i.e. the data presented are the results of biological replicates. chemicals the pua used in this study was 2e,4e/z-octadienal (octa, from sigma aldrich inc. milan, italy). this was chosen because it is the main and most commonly produced pua by diatoms, and, in particular, by skeletonema spp. (wichard et al., 2005). the concentrations used were chosen accounting for multiples of the concentration causing a reduction of 50% in cell concentrations at 24h, as empirically determined for s. tropicum (data not shown) resulting to be 8 μm. the reference value of 8 μm has been chosen also based on ribalet et al., (2007), where 8 μm represented the ec50 for growth in the congeneric species s. marinoi. octa working solutions were dissolved in hplc-grade pure methanol (jt baker, phillipsburg, nj, usa) at room temperature. since methanol can be toxic, it was tested on growing cultures and cultures showed no reaction up to 7 microliters per ml. subsequently, and conservatively, no more than 4 microliters of methanol plus octa were ever inoculated into the cultures. before incubation, the exact octa concentration in the working solution was assessed spectrophotometrically (hewlett-packard 8453hewlett-packard co., palo alto, ca, usa) by measuring absorption at 274 nm and a specific molar absorption coefficient of 31,000 (pippen and nonaka, 1958). hydrogen peroxide (h2o2, 30 wt. % from sigma aldrich, milan, italy) was used as positive control, after dilution 1/200 in filtered seawater and added to samples as a 50 μl:1 ml ratio. the ros scavenger, 4-hydroxy-2,2,6,6-tetramethylpiperidine 1-oxyl (tempol, sigma-aldrich, milan, italy) was used as negative control, so as sodium diethyldithiocarbamate trihydrate (detc, sigma-aldrich), an inhibitor of superoxide dismutase (sod) at a final concentration of 5 mm and 1 mm, respectively, in filtered sea water (fsw). samples with detc were incubated for 120 min prior to pua addition. the ros-sensitive dye dihydrorodamine 123 (dhr, molecular probes, leiden, nl) (5 mm stock in dmso) was used to monitor ros production at the single-cell level. dhr can be oxidized by different ros, including hydrogen peroxide (h2o2) and peroxynitrite (onoo–), to form the fluorescent derivative rhodamine 123 (excitation wl 492-495 nm, emission wl=517-527 nm), and is commonly used to investigate oxidative stress in vivo (ischiropoulos et al., 1999). dhr has been previously used in different microalgal species to measure oxidative burst (jamers et al., 2009; vardi et al., 1999). samples were incubated with both 10 μm dhr and pua in the dark at room temperature and analysed using flow cytometry (see below). flow cytometry to quantify ros production in vivo, a flow cytometer (facscalibur, becton dickinson, palo alto, ca, usa) equipped with a 488-nm laser as excitation source and filtered sea water as sheath was used. a 530/30-nm beampass emission filter was used for detection of dhr-derived green fluorescence. both the sheath fluid and the sample velocity (65 μlmin–1) were kept constant during all the experiments. as a trigger signal, red fluorescence was used with a threshold at channel 52. red fluorescence was collected through a 650-nm long-pass filter and was also used as a proxy for chlorophyll cell fluorescence. fluorescent beads (coulter flowset fluorospheres, beckman coulter, fullerton, ca, usa) were used as internal standard. green fluorescence values were normalized to the green fluorescence of the beads, and all data were expressed as fold changes of green fluorescence relative to the control values. data acquisition (104 cells on average for each sample) was performed using cellquest software (becton-dikinson). data analysis was performed using fcs4 express (de novo softwares, los angeles, ca, usa). total rna extraction two-hundred ml of culture were collected through a filtration system using a 2-µm pore size polycarbonate filter (nucleopore; whatman, maidstone, uk), placed over a no nco mm er cia l u se on ly 16dsp expression in response to pua gf/f glass microfiber filter (whatman). filtration never lasted longer than 30 min. cells were then resuspended in 0.7 ml guanidine isothiocyanate buffer (rlt buffer, qiagen, valencia, ca, usa) containing 1% β-mercaptoethanol (sigma aldrich, st. louis, mo, usa), immediately frozen at -80°c, and stored for a few days. after disrupting the cells by sonication on ice (sonicator ultrasonic processor xl, heat system ultrasonics, farmingdale, ny, usa), total rna was extracted using the silica-membrane spin column included in the rneasy plant mini kit (qiagen). the residual genomic dna was removed by an additional treatment with the rnase-free dnase set (qiagen) according to the manufacturer’s instruction. total rna concentration and purity were determined with a spectrophotometer (nd-100; nanodrop technologies, wilmington, de, usa) by measuring absorption at 260 and 280 nm. reverse transcription and real time quantitative polymerase chain reaction one μg of total rna was treated with dnase i and reverse transcribed into first-strand complementary dna (cdna) fragments by using random hexamer oligonucleotides and the high capacity cdna reverse transcription kit (applied biosystems, foster city, ca, usa). the conditions for the retro-transcription were set as in chung et al. (2005): 25°c for 10 min, 37°c for 120 min, and 85°c for 5 min. real time quantitative polymerase chain reactions (q-pcr) were performed using 2× sybr green pcr master mix (applied biosystems) with 250 nm of each of the forward and reverse primers. the nucleotide sequences of all primer pairs used in the q-pcrs are indicated in tab. 1. the geneamp 7000 sequence detection system (applied biosystems) was used to carry out all the reactions. the threshold cycle at which the fluorescence intensity exceeded a preset threshold was used to calculate the target gene mrna and 18s rrna expression levels. the rna molar ratio of scdsp mrna and 18s rrna was calculated using the formula described in chung et al., (2005), and the data normalized to the control values. the levels of scdsp mrna expression were expressed as fold changes of mrna abundance relative to the control. all data reported are means of three technical replicates but the experiments were repeated at least twice. statistical analyses pairwise comparisons among treatments and control were assessed by student’s t-test using the excel spreadsheet (microsoft, redmond, washington, ma, usa). results and discussion cultures were affected by octa exposure. 16 μm octa reduced cell concentrations to 51.17% of initial values after 24 h (fig. 1; tab.2). h2o2 also strongly affected cell concentrations which decreased to 50% as early as 2 h after inoculation and to zero 24 h later (fig. 1). scdsp-1 and scdsp-2 gene expression and ros levels increased in s. tropicum cultures exposed to different concentrations of octa. a peak in scdsp-1 and scdsp2 expression was visible as early as 1 h after exposure and its intensity for cells in le exposed to the concentration of 8 μm octa was ca. 822 and 98 folds that in the control for scdsp-1 and scdsp-2, respectively (fig. 2 a,b). gene expression decreased afterward and showed no further peak until the end of the experiment. in the same strain, an increase in scdsp expression has been previously reported when cultures entered senescence (chung et al., 2005) and also in response to a light stress involving the blockage of electron flow at the level of the photosystem ii (chung et al., 2008). the latter also involved no increase in the gene regulation and involvement of a pcd machinery (chung et al., 2008). overall, these results provide evidence that the dsp genes have a role in regulating the molecular mechanism of mortality in s. tropicum under stress. in our experiments, gene expression was highest at the concentrations where ros production was intermediate, suggesting that there exists a threshold for the activation of the two death-related genes investigated. a peak in ros was observed 20 min after inoculation with octa (fig. 3), followed by a decrease. the intensity of the peak tab. 1. primer pairs used for the qrt pcr reactions, as in chung et al. (2008). target gene primer name nucleotide sequence (5’-end to 3’-end) scdsp-1 scdsp-sg-f gaaca agcaa actgc actcg tc scdsp-sg-r gtcaa gaatg ttggt cgtcg cg scdsp-2 scdsp-sg-f gaaca agcaa actgc actcg tc scdsp2-sg-r gtagg catct gctat tcttt ctg 18s rrna ske-18s-f gaatt cctag atatc gcagt tcatc ske-18s-r gctaa tccac aatct cgact cctc tab. 2. effect of octa on growth rate of s. tropicum cultures after 24 h. initial concentration (t0, i.e., le) was 2.53 105 cell ml–1. growth rate (day–1) concentration (µm) le octa° 0 0.30±0.10 4 0.03±0.08 8* -0.63±0.39 16 -0.80±0.22 16h2o2 2.70±0.35 le, late exponential growth phase (as in fig. 1); °2e,4e/z-octadienal; *concentration causing a reduction of 50% in cell concentrations at 24 h (see text). data are means ±sd, n=3. no nco mm er cia l u se on ly 17 a.a. gallina et al. was pua concentration-dependent, and was highest for 16 μm octa (1.9±0.2 increase; p<0.001, n=3). both the ros scavenger tempol and the sod inhibitor detc reduced the intensity of the peak, therefore supporting the conclusion that the peak was really due to an increase of ros. no increase in dhr-derived green fluorescence was observed when methanol only was added to the cultures, excluding an artifact due to the solvent used (data not shown). hence, it is suggested that cell death is triggered in s. tropicum only by specific pua concentrations, which are those causing a reduction of 50% in cell concentrations at 24 h (8 μm octa). this is supported by the observation that 8 μm octa induced intermediate levels of ros production (1.5±0.1 increase in le cultures), while a higher concentration (16 μm octa) elicited a considerably lower scdsp expression, even though it induced a higher ros production (1.9±0.2 increase). this could be related to different molecular signals triggered in the two cases. in our experiments, while the addition of h2o2 to the cultures induced an increased dsp gene expression, the levels were not as high as those elicited by octa, even though the dhr-ros derived green fluorescence was the highest observed (7.5±0.9 times the control at 40 min after exposure, not shown). this suggests that ros could be important modulators in the pathway leading to cell death. even though evidence for the occurrence of apoptotic-like features in s. tropicum in response to pua was not provided in this study, it has been previously reported that pua are able to induce apoptotic-like cell death in the marine diatom thalassiosira weissflogii (casotti et al., 2005). in addition, in the marine diatom phaeodactylum tricornutum pua exposure induced overexpression of a gene associated to no production (designated as ptnoa), that resulted in an altered expression of superoxide dismutase and an increased metacaspase activity, which are known factors in the stress responses and pcd pathway (vardi et al., 2008). indeed, in our octa-exposed s. tropicum cultures, pua had a toxic effect drastically affecting growth and inducing cell death. ros are generally indicated in the literature as pcdinducing factors (aldsworth et al., 1999; simon et al., 2000). however, we suggest that a difference exists between the intracellular effectors of a pcd-like process (e.g., no and ros), and other factors and stimuli that, causing an alteration of no and ros, are able to modulate the response leading to cell death. based on the results reported here, it is proposed that pua act as such factors, which induce modifications of the intracellular ros levels, which then result in a differential activation of scdsp genes. the successive reaction depends then on the amount of ros produced and possibly on the cell physiological and growth status. this scenario is supported by the observation that the increase in dsp gene expression in our experiments was not linearly correlated to pua concentrations. while it is possible that the experimental design was not appropriate to detect sudden peaks occurring at intermediate time points, it is unlikely that these lasted for shorter than 1 h. indeed, when a further time point at 30 min after inoculation was included, the expression levels were not even comparable to the highest expression measured during the experiments (data not shown). therefore, fig. 2. scdsp1 (a) and scdsp2 (b) mrna expression in s. tropicum cultures exposed to different octa concentrations and h2o2 in le cultures. data are expressed as mean of fold changes of mrna abundance (relative to control) from technical triplicates. the experiments were repeated at least twice. no nco mm er cia l u se on ly 18dsp expression in response to pua considering that a burst in ros production always occurred 20 min after exposure to all the different pua tested, it is possible to conclude that the concentration of 8 μm octa represents a threshold value below which the cell activates defences mechanisms to cope with the related oxidative stress. (chung et al., 2005, 2008). based on the results reported here, it is suggested that in the case of pua-induced chemical stress, dsp genes expression is regulated by ros. although we cannot exclude a role for no in the process, data from the congeneric s. marinoi show that ros and not no have a main role in the response to pua in this diatom and probably in the congeneric s. tropicum (gallina et al., 2014). in addition, considered that dsp is reported to be a ca2+ regulated protein (chung et al., 2005), it is then possible that in s. tropicum ros are acting as ca2+ mobilizing agents, as also happens in other biological systems (gonzalez et al., 2002; mallilankaraman et al., 2011; wendehenne et al., 2004). conclusions the reported findings are only preliminary and need to be substantiated by further experiments aimed at proving a clear connection between ros and dsp expression. in addition, direct data on pcd activation have to be acquired, possibly using the genomic approach used by lauritano et al., (2015). however, based on these preliminary results and combining them on the vast literature on pua in diatoms, we propose that in skeletonema spp. the ros downstream response to pua activates either a protection pathway (i.e., antioxidant defence) or a cell death cascade either through activation of specific genes or through nongenetically controlled cell death, depending on the pua used, its concentration, the time of exposure, the concentration of ros produced and the physiological state of the cell. this has important implications for bloom regulation and population dynamics and induction of cell lysis, as instance, immediately after the peak stage of a diatom bloom, when cell concentrations are highest but environmental factors become limiting and do not support further growth (vanboekel et al., 1992). it has been proposed that the evolution of pcd in unicellular organisms as a byproduct of selection might be explained by the possibility that pcd could be a maladaptive stress-induced process that can be selected under certain conditions to benefit only some organisms or part of a population (nedelcu et al., 2010). as related to this concept, if an intrapopulation signal (e.g., pua) triggers the production of pcd-inducing signal (e.g., ros), and if this intrapopulation signal-dependent death is beneficial for a part of the population (for instance, by fitting cell density to the environmental conditions), then pcd could be adaptively selected for and consequently evolve into altruistic adaptation (nedelcu et al., 2010). this hypothesis needs to be tested in situ, where pcd has been shown to exist, for instance in reaction to a viral attack (bidle et al. 2007), supporting its role as population regulatory mechanism, or, triggering the production of survival forms such as cysts or resting stages, using high ros levels as signals (vardi et al., 1999), so to assure perpetuation of the species later on, when the environment conditions become favourable again (mcquoid et al., 2002). acknowledgments aag deeply acknowledges sh hung and sf. wang for their assistance during experiments and technical support. travel fund for aag has been supported by the national science council of taiwan (nsc) through the summer program for italian graduate students fellowship. aag has been supported by a szn-ou phd fellowship. materials for this study were partly supported by the taiwan national science council (nsc 99-2611-m-019 -014 -my2). references aldsworth tg, sharman rl, dodd cer, 1999. bacterial suicide through stress. cell. mol. life sci. 56: 78-383. andrianasolo eh, haramaty l, degenhardt k, mathew r, white e, lutz r, falkowski p, 2007. induction of apoptosis by diterpenes from the soft coral xenia elongata. j. nat. prod. 70:1551-1557. fig. 3. ros production in s. tropicum exposed to different concentrations of octa. data are values of dhr123-derived green fluorescence normalized to the control values. data are means ±sd from three biological replicates. no nco mm er cia l u se on ly 19 a.a. gallina et al. armbrust ev, 2009. the life of diatoms in the world’s oceans. nature 459:185-192. berges ja, falkowski pg, 1998. physiological stress and cell death in marine phytoplankton: induction of proteases in response to nitrogen or light limitation. limnol. oceanogr. 43:129-135. bidle kd, bender sj, 2008. iron starvation and culture age activate metacaspases and programmed cell death in the marine diatom thalassiosira pseudonana. eukaryot. cell 7:223-236. bidle kd, falkowski pg, 2004. cell death in planktonic, photosynthetic microorganisms. nat. rev. microbiol. 2:643-655. bidle kd, haramaty l, ramos jbe, falkowski p, 2007. viral activation and recruitment of metacaspases in the unicellular coccolithophore, emiliania huxleyi. p. natl. acad. sci. usa 104:6049-6054. butow bj, wynne d, telor e, 1997. antioxidative protection of peridinium gatunense in lake kinneret: seasonal and daily variation. j. phycol. 33:780-786. casotti r, mazza s, brunet c, vantrepotte v, ianora a, miralto a, 2005. growth inhibition and toxicity of the diatom aldehyde 2-trans, 4-trans-decadienal on thalassiosira weissflogii (bacillariophyceae). j. phycol. 41:7-20. cervia d, di giuseppe g, ristori c, martini d, gambellini g, bagnoli p, dini f, 2009. the secondary metabolite euplotin c induces apoptosis-like death in the marine ciliated protist euplotes vannus. j. eukaryot. microbiol. 56:263-269. chandra j, samali a, orrenius s, 2000. triggering and modulation of apoptosis by oxidative stress. free radical biology and medicine 29, 323-333. chung cc, hwang spl, chang j, 2005. cooccurrence of scdsp gene expression, cell death, and dna fragmentation in a marine diatom, skeletonema costatum. appl. environ. microbiol. 71:8744-8751. chung cc, hwang spl, chang j, 2008. nitric oxide as a signaling factor to upregulate the death-specific protein in a marine diatom, skeletonema costatum, during blockage of electron flow in photosynthesis. appl. environ. microbiol. 74:6521-6527. costas e, aguilera a, gonzalezgil s, lopezrodas v, 1993. contact inhibition: also a control for cell proliferation in unicellular algae? biol. bull. 184:1-5. deponte m, 2008. programmed cell death in protists. bba mol. cell res. 1783:1396-1405. ellegaard m, ribeiro s, lundholm n, andersen tj, berge t, ekelund f, harnstrom k, godhe a, 2013. using the sediment archive of living dinoflagellate cysts and other protists resting stages to study dynamics through time, pp. 149-153. in: f. marrett, j.m. lewis and l.r. bradely (eds.), biological and geological perspectives of dinoflagellates. geological society, london. evans c, malin g, mills gp, wilson wh, 2006. viral infection of emiliania huxleyi (prymnesiophyceae) leads to elevated production of reactive oxygen species. j. phycol. 42:1040-1047. gallina aa, brunet c, palumbo a, casotti r, 2014. the effect of polyunsaturated aldehydes on skeletonema marinoi (bacillariophyceae): the involvement of reactive oxygen species and nitric oxide. mar. drugs 12:4165-4187. gonzalez a, schmid a, salido gm, camello pj, pariente ja, 2002. xod-catalyzed ros generation mobilizes calcium from intracellular stores in mouse pancreatic acinar cells. cell. signal. 14:153-159. guillard rrl, 1975. culture of phytoplankton for feeding marine invertebrates, pp. 29-60. in:, w. l. smith and m.h. chanley (eds.), cultures of marine invertebrate animals. plenum publishing co., new york. ischiropoulos h, gow a, thom sr, kooy nw, royall ja, crow jp, 1999. detection of reactive nitrogen species using 2,7dichlorodihydrofluorescein and dihydrorhodamine 123. in: j.n. abelson, m.i. simon and h. sies (eds.). methods in enzymology nitric oxide part c biological and antioxidant activities, vol. 301. academic press, san diego. jamers a, lenjou m, deraedt p, van bockstaele d, blust r, de coen w, 2009. flow cytometric analysis of the cadmiumexposed green alga chlamydomonas reinhardtii (chlorophyceae). eur. j. phycol. 44, 541-550. jimenez c, capasso jm, edelstein cl, rivard cj, lucia s, breusegem s, berl t, segovia m, 2009. different ways to die: cell death modes of the unicellular chlorophyte dunaliella viridis exposed to various environmental stresses are mediated by the caspase-like activity devdase. j. exp. bot. 60:815-828. jones dp, 2008. radical-free biology of oxidative stress. am. j. physiol. cell physiol. 295:c849-c868. kuwabara m, asanuma t, niwa k, inanami o, 2008. regulation of cell survival and death signals induced by oxidative stress. j. clin. biochem. nutr. 43:51-57. lam e, 2008. programmed cell death in plants: orchestrating an intrinsic suicide program within walls. crit. rev. plant sci. 27:413-423. lauritano c, orefice i, procaccini g, romano g, ianora a, 2015. key genes as stress indicators in the ubiquitous diatom skeletonema marinoi. bmc genomics 16:411. luan s, kudla j, rodriguez-concepcion m, yalovsky s, gruissem w, 2002. calmodulins and calcineurin b-like proteins: calcium sensors for specific signal response coupling in plants. plant cell 14:s389-s400. mallilankaraman k, gandhirajan rk, hawkins bj, madesh m, 2011. visualization of vascular ca2+ signaling triggered by paracrine derived ros. j. vis. exp. doi: 10.3791/3511 martindale jl, holbrook nj, 2002. cellular response to oxidative stress: signaling for suicide and survival. j. cell. physiol. 192:1-15. mcquoid mr, godhe a, nordberg k, 2002. viability of phytoplankton resting stages in the sediments of a coastal swedish fjord. eur. j. phycol. 37:191-201. means ar, dedman jr, 1980. calmodulin an intracellular calcium receptor. nature 285:73-77. nedelcu am, driscoll ww, durand pm, herron md, rashidi a, 2010. on the paradigm of altruistic suicide in the unicellular world. evolution 65:3-20. pippen el, nonaka m, 1958. a convenient method for synthesizing normal aliphatic 2,4-dienals. j. org. chem. 23:15801582. ribalet f, berges ja, ianora a, casotti r, 2007. growth inhibition of cultured marine. phytoplankton by toxic algal-derived polyunsaturated aldehydes. aquat. toxicol. 85:219-227. rockenfeller p, madeo f, 2008. apoptotic death of ageing yeast. exp. gerontol. 43: 876-881. scherz-shouval r, elazar z, 2007. ros, mitochondria and the no nco mm er cia l u se on ly 20dsp expression in response to pua regulation of autophagy. trends cell biol. 17:422-427. segovia m, 2011. cell death or survival in the unicellular chlorophyte dunaliella: the role of metacaspases and caspase-like proteases. eur. j. phycol. 46:32-33. segovia m, berges ja, 2009. inhibition of caspase-like activities prevents the appearance of reactive oxygen species and dark-induced apoptosis in the unicellular chlorophyte dunaliella tertiolecta. j. phycol. 45:1116-1126. simon hu, haj-yehia a, levi-schaffer f, 2000. role of reactive oxygen species (ros) in apoptosis induction. apoptosis 5:415-418. thamatrakoln k, bailleul b, brown cm, gorbunov my, kustka ab, frada m, joliot pa, falkowski pg, bidle kd, 2013. death-specific protein in a marine diatom regulates photosynthetic responses to iron and light availability. p. natl. acad. sci. usa 110:20123-20128. thamatrakoln k, korenovska o, niheu ak, bidle kd, 2012. whole-genome expression analysis reveals a role for deathrelated genes in stress acclimation of the diatom thalassiosira pseudonana. environ. microbiol. 14:67-81. thibane vs, ells r, hugo a, albertyn j, van rensburg wjj, van wyk pwj, kock jlf, pohl ch, 2012. polyunsaturated fatty acids cause apoptosis in c. albicans and c. dubliniensis biofilms. biochim. biophys. acta 1820.1463-1468. vanboekel whm, hansen fc, riegman r, bak rpm, 1992. lysis-induced decline of a phaeocystis spring bloom and coupling with the microbial food web. mar. ecol.-progr. ser. 81:269-276. vardi a, berman-frank i, rozenberg t, hadas o, kaplan a, levine a, 1999. programmed cell death of the dinoflagellate peridinium gatunense is mediated by co2 limitation and oxidative stress. curr. biol. 9:1061-1064. vardi a, bidie kd, kwityn c, hirsh dj, thompson sm, callow ja, falkowski p, bowler c, 2008. a diatom gene regulating nitric-oxide signaling and susceptibility to diatom-derived aldehydes. curr. biol. 18:895-899. vardi a, eisenstadt d, murik o, berman-frank i, zohary t, levine a, kaplan a, 2007. synchronization of cell death in a dinoflagellate population is mediated by an excreted thiol protease. environ. microbiol. 9:360-369. vardi a, formiggini f, casotti r, de martino a, ribalet f, miralto a, bowler c, 2006. a stress surveillance system based on calcium and nitric oxide in marine diatoms. plos biol. 4:411-419. vardi a, schatz d, beeri k, motro u, sukenik a, levine a, kaplan a, 2002. dinoflagellate-cyanobacterium communication may determine the composition of phytoplankton assemblage in a mesotrophic lake. curr. biol. 12:1767-1772. vardi a, van mooy bas, fredricks hf, popendorf kj, ossolinski je, haramaty l, bidle kd, 2009. viral glycosphingolipids induce lytic infection and cell death in marine phytoplankton. science 326:861-865. wendehenne d, durner j, klessig df, 2004. nitric oxide: a new player in plant signalling and defence responses. curr. opin. plant biol. 7:449-455. wichard t, poulet sa, halsband-lenk c, albaina a, harris r, liu dy, pohnert g, 2005. survey of the chemical defence potential of diatoms: screening of fifty one species for alpha,beta,gamma,delta-unsaturated aldehydes. j. chem. ecol. 31:949-958. no nco mm er cia l u se on ly layout 1 introduction cyanobacterial blooms cause local and global problems by contaminating surface water resources with their diverse types of potent toxins commonly known as cyanobacterial toxins or cyanotoxins. the most frequently reported and widespread cyanotoxins are the cyclic heptapeptide microcystins (molecular weight ~1000 da) in fresh water and penta-peptide nodularins (molecular weight 825 da) in brackish water. microcystins and nodularins are potent hepatotoxins (liver toxins) with an acute ld50 value of 25 to ~1000 µg kg–1 (mouse, i.p.) (codd et al., 2005). besides acute toxicity, microcystins and nodularins are tumor promoters (nishiwaki-matsushima et al., 1992; sueoka et al., 1997) and microcystin-lr is described as a possible carcinogen (grosse et al., 2006). microcystins are mainly produced by the most common cyanobacteria genera found world-wide: microcystis, dolichospermum, nostoc, planktothrix, anabaenopsis, and hapalosiphon. the brackish water cyanobacterium nodularia is the main producer of nodularins (codd et al., 2005). overall occurrences of these toxins in surface water resources pose detrimental health threat to human and various animals including livestock, wild mammals and birds (codd et al., 2005; stewart et al., 2008; merel et al., 2013). the world health organization guideline limit for microcystin-lr in drinking water is 1 µg l–1 (who, 2011). simple and efficient methods for cyanotoxin detection are in high demand in order to assess the quality of water sources used for drinking water abstraction and for recreational or agricultural use. however, the structural diversity of this toxin family constitutes a great challenge when monitoring water or making an assay. the unusual β-amino acid adda (3-amino-9-methoxy-2,6,8-trimethyl-10phenyldeca-4(e),6(e)-dienoic acid) (botes et al., 1984; rinehart et al., 1988) is common for both microcystins and nodularins; and has important role in toxicity (dahlem, 1989). other structural features necessary for the toxicity include the cyclic structure of the toxin (choi et al., 1993; rinehart et al., 1994) as well as the free carboxylic acid group in the d-glu unit (stotts et al., 1993). a change of advances in oceanography and limnology, 2017; 8(1): 121-130 article doi: 10.4081/aiol.2017.6349 this work is licensed under a creative commons attribution-noncommercial 4.0 international license (cc by-nc 4.0). non-competitive elisa with broad specificity for microcystins and nodularins sultana akter,1* markus vehniäinen,1 jussi meriluoto,2 lisa spoof,2 urpo lamminmäki1 1molecular biotechnology and diagnostics, department of biochemistry, university of turku, 20520 turku; 2biochemistry, faculty of science and engineering, åbo akademi university, fi-20520 turku, finland *corresponding author: sultana.akter@utu.fi abstract simple and cost-effective methods with sufficient sensitivities for preliminary screening of cyanobacterial toxins are in high demand for assessing water quality and safety. we have recently developed a highly sensitive and rapid time-resolved fluorometry based noncompetitive immunoassay for detection of microcystins and nodularins. the assay is based on a synthetic broad-specific anti-immunocomplex antibody sa51d1 capable of recognizing the immunocomplex formed by a generic anti-adda monoclonal antibody (mab) bound to either microcystins or nodularins. using the same antibody pair, here we describe a very simple and cost-efficient non-competitive elisa test for microcystins and nodularins based on conventional alkaline phosphatase (ap) activity measurement. the recombinant sa51d1 single-chain fragment of antibody variable domain (scfv) was produced as a fusion with bacterial alkaline phosphatase in escherichia coli. after one step affinity purification through his-tag, the scfv-ap fusion protein could directly be used in the assay. for the assay, toxin standard/sample, biotinylated anti-adda mab and the scfv-ap were incubated together for one hour on streptavidin-coated microtiter wells, washed and ap activity was then measured by incubating (1 h at 37°c) with chromogenic substrate para-nitrophenylphosphate (pnpp). the assay was capable of detecting all the eleven tested toxin variants (microcystin-lr, -dmlr, -rr, -dmrr, -yr, la ly, -lf -lw, -wr, and nodularin-r) below who guide line value of 1 µg l–1. the detection limit (based on blank+3sd response) for microcystin-lr was 0.2 µg l–1. the assay was verified using spiked (0.25-4 µg l–1 of microcystin-lr) tap, river and lake water samples with recoveries from 64 to 101%. the assay showed good correlation (r2>0.9) with four reference methods for its performance in detecting extracted intracellular microcystin/nodularin from 17 natural surface water samples. the described easy-to-perform assay has a high potential to be used in resource-poor settings as quantitative measurements can be obtained using a simple elisa reader or easy-to-interpret qualitative results by visual readout. based on the non-competitive format, the assay does not need any chemical toxin conjugates and offers robustness as compared to the currently available competitive format assays. key words: immunocomplex assay; sandwich-type elisa; cyanotoxin; microcystin; nodularin; harmful algal bloom. received: 17 october 2016. accepted: 31 january 2017. no nco mm er cia l u se on ly 122 non-competitive broad specificity elisa for cyanotoxins two (microcystin) or one (nodularin) amino acid(s), together with several other structural modifications such as methylation, hydroxylation, and epimerization in small side groups during nonribosomal peptide synthesis creates diversity (rinehart et al., 1994; sivonen, 1996; neilan et al., 1999).variation has been observed in all positions of microcystins and nodularins; and close to 250 analogues of microcystin and 10 analogues of nodularin with differing toxicity have been reported to date in the literature (sivonen and jones, 1999; mazur-marzec et al., 2006; puddick, 2013; niedermeyer, 2014; spoof and catherine, 2017). different laboratory analysis methods exist for measuring microcystins and nodularins from water bodies and environmental samples. in general, immunoassays (commonly known as elisa) or protein phosphatase inhibition assay (ppia) are used for preliminary screening, followed by quantification and identification by sophisticated methods like high-performance liquid chromatography (hplc) coupled to mass spectrometry (ms) for positive samples (meriluoto and codd, 2005; lawton and edwards, 2008). immunoassay techniques are emerging popular methods due to their simplicity together with easiness to handle. water samples (either tap water or from natural water bodies) can be analyzed as such with immunoassay techniques. immunoassays are amenable for automation and screening of large number of samples and they can reduce the need for more expensive and time consuming analyses (sivonen, 2008). though the direct non-competitive immunoassay offers theoretical advantages (for example, in terms of sensitivity, high specificity, flexibility etc.) over the competitive format assays, all the currently available commercial immunoassay kits for microcystins/nodularins are in the competitive format. the main reason is that the molecular weight of cyanotoxin is about 1000 daltons, making it difficult or even impossible to have two independent binding sites on its surface needed for non-competitive immunoassays. non-competitive assays rely on direct measurement of antibody binding sites occupied by analyte producing an easy-to-interpret signal which is proportional to analyte concentration. on the other hand, competitive assays rely on measurement of unoccupied sites, from which the occupied sites are inferred by subtraction (deshpande, 1996; self et al., 2013). due to the indirect approach the competitive assays usually require several assay steps and produce decreasing signal with increasing analyte concentration. as the signal in the absence of the analyte is already high, it is difficult to discern the slight changes in the signal caused by low analyte concentrations, especially with visual detection. also, the competitive assays usually require strict maintaining of the reaction condition since equilibrium must be reached (deshpande, 1996; self et al., 2013). we have recently reported isolation of a unique generic anti-immunocomplex binder from our in-house synthetic antibody library and development of a broad-spectrum noncompetitive immunocomplex immunoassay for microcystins and nodularins (akter et al., 2016). the described time-resolved fluorometry (trf) based assay is highly sensitive and rapid. however, in addition to a lanthanide chelate (europium) labeled tracer reagent, the assay requires a fluorometer with trf detection capability which is unfortunately rarely found in laboratories. in order to omit the need of special reagent and instrumentation, we describe here an elisa method based on the aforementioned immunocomplex assay principle using particularly simple-toproduce components and easily accessible detection chemistry. the capability of the assay for broad-spectrum detection of microcystins and nodularins is demonstrated using purified toxins and environmental samples that were also tested with several reference methods. methods common materials and instruments common inorganic and organic chemical reagents were obtained from commercial source either from sigma or merck unless otherwise specified. the reagent water used was purified by millipore milli-q plus water filtration purification system (millipore corporation, bedford, ma, usa). multilabel counter victor 1420 for signal measurement was from wallac/perkinelmer life sciences, (waltham, ma, usa). streptavidin coated microtiter plates were from kaivogen oy (turku, finland). monoclonal antibody, ad4g2 (adda specific, anti-microcystins) was from enzo life sciences, inc. (farmingdale, ny, usa) and was biotinylated with biotinisothiocyanat (bitc) to be captured on streptavidin surface of microtiter well. assay buffer was composed of 50 mm tsa (tris saline azide buffer) ph 7.75 supplemented with 0.01% tween 40, 0.05% bovine-γ-globulin, 20 µm dtpa (diethylenetriaminepentaacetic acid), 0.5% bovine serum albumin (bsa) and finally (optional) 20 µg ml–1 of amaranth dye solution (cas: 915-67-3, sigma) to aid pipetting. wash buffer for washing of microtiter well-plate contained 5 mm tris-hcl ph 7.75, 0.9% nacl, 0.1% germall ii, and 0.005% tween 20. the bacterial host escherichia coli (e. coli) xl1-blue was from agilent technologies (santa clara, ca, usa) and used for expression of the singlechain fragment of antibody variable domain (scfv). paranitrophenyl phosphate (pnpp) substrate 4-nitrophenyl phosphate disodium salt hexahydrate was from sigmaaldrich (st. louis, mo, usa). toxin standards specific amount of the purified toxins (supplementary fig. 1) (microcystin-lr, -dmlr, -rr, -dmrr, -yr, -la -ly, -lf -lw, and nodularin-r) were obtained from dr. meriluoto’s lab (åbo akademi university) as a no nco mm er cia l u se on ly 123s. akter et al. lyophilized dried powder. the toxins were purified by preparative hplc according to methods described earlier (meriluoto and codd, 2005). microcystin-la and microcystin-wr were purchased from enzo life science. all the toxin standards were stored dry at -20°c until required. dry powder was dissolved in 50% methanol (100250 µg ml–1 original stock solution) and kept at -20°c or 4°c in sealed condition. from these original stocks, further working solutions and standards were prepared in reagent water and stored at -20°c, or at 4°c. the maximum percentage volume of methanol in the toxin standard solution (0.02-600 µg l–1) used in the assay was 0.3%. generic anti-immunocomplex binder for microcystin and nodularin the clone sa51d1scfv-ap (generic anti-immunocomplex binder for microcystin and nodularin) used in this study does not show detectable specificity to the naked antiadda monoclonal antibody (mab) or to the toxin alone and is capable of recognizing the immunocomplexes composed of anti-adda mab bound to any of the eleven tested cyanotoxin analogues (supplementary fig. 1). the isolation and characterization of the binder antibody was described in detail earlier (akter et al., 2016). production and purification of scfv-ap fragments e. coli cells carrying the clone sa51d1 construct in plk06h (huovinen et al., 2013) vector were grown in 50 ml shaking flask in sb medium supplemented with 100 µg ml–1 ampicillin, 10 µg ml–1 tetracycline and 0.05% glucose. the cells were induced with iptg (isopropyl-βd-1-thiogalactopyranoside induction) to the final concentration of 100 µm and incubated overnight at 26°c, with shaking at 300 rpm. cells from the culture supernatant were used to purify the scfv-ap protein by his affinity column (his spin trap™ kit, ge healthcare, uk) according to the manufacturer’s instructions. non-competitive elisa one hundred µl per well of reagent water (for blank measurement, 6-56 replicates per assay), toxin standard solution (of concentration: 0.02-600 µgl–1, prepared in reagent water), or samples were added in prewashed streptavidin wells (2-4 replicates for standard or sample). then, 100 µl per well of antibodies (biotinylated anti-adda mab, 1 µg ml–1 and scfv-ap, 1 µg ml–1) prepared in assay buffer were added to each well. the wells were incubated (with slow shake) for 1 h at room temperature (rt, ~23°c) followed by four washes. then pnpp liquid substrate solution (5 mm) prepared in 0.1 m glycine buffer, 1 mm mgcl2, 1 mm zncl2, ph 10.4 was added (200 µl per well). plates were incubated at 37°c for 1 h and absorbance was measured at 405 nm. the assay concept and procedure is illustrated in supplementary fig. 2. sample concentrations were calculated from the standard curve (microcystin-lr) using origin 2015 software (originlab corporation, wellesley hills, usa). the detection limit (the smallest detectable toxin concentration in sample) was calculated from the standard curve based on the average response of “n” (n=6 to 56) number of replicates of blank plus 3 times standard deviation of the blank. influence of temperature and incubation time on assay performance the effect of temperature and incubation time on the assay performance during initial bioaffinity reaction and during the signal development stage (ap activity) was tested using microcystin-lr standard of concentration 0.02600 µg l–1. the assay protocol was same as described above with following exception. the initial bioaffinity reaction was tested either at rt or at 37°c for 30 min and 1 h. after the washing step, pnpp liquid substrate was added and incubated either at rt or at 37°c. the measurements were carried out at different time points (30 min to 24 h). non-competitive elisa with different toxin variants the standard curves (toxin standard concentration: 0.02 to 600 µg l–1) of eleven microcystin/nodularin analogues (supplementary fig. 1) were obtained (duplicate measurements, for blank, n=56) using the protocol described above. signal development after washing step was done at 37°c; absorbance was measured after 1 h. non-competitive elisa with spiked water samples a total of four water samples including one reagent water, one tap water sample from our laboratory, one river water (paimio river) and one lake water (paalijärvi) samples from finland were spiked with microcystin-lr at concentration from 0.25 to 4 µg l–1 (tab.1). the environmental surface water samples were collected during 2009 (tab. 2) and were stored as such at -20°c until use. upon thawing at rt, samples were spiked with microcystin-lr over a range of concentrations (0, 0.25, 0.5, 1, 2 and 4 μg l–1). the spiked samples and the corresponding unspiked samples were measured by the noncompetitive elisa in duplicates. the unspiked samples were also measured by a commercial immunoassay (microcystins-adda elisa, abraxis, pa, usa) and by the time resolved fluorescent measurement based non-competitive immunoassay (akter et al., 2016). presence of toxin in the unspiked samples by any of the three tested methods was taken into consideration for recovery calculation. the recovery percentage of the spiked sample was calculated as follows: r%=(spiked sample result unspiked sample result) x (known spike added concentration)–1 x 100%. no nco mm er cia l u se on ly 124 non-competitive broad specificity elisa for cyanotoxins non-competitive elisa with surface environmental water samples a total of seventeen environmental surface water samples from a sample panel collected during 2009 from finland and estonia were tested using the non-competitive elisa for internal (cellular) toxin in water. the samples constituted of cyanobacterial cells harvested on filters which were extracted for intracellular toxins with 75% methanol by method described earlier (hautala et al., 2013; savela et al., 2014). the methanolic extracts were aliquoted (100-500 μl extract) and the aliquots evaporated dry. the individual aliquots were re-dissolved in reagent water for ppia and in 75% methanol for hplc. the ppia method was based on protein phosphatase 1 inhibition of p-nitrophenyl phosphate cleavage followed spectrophotometry (using microcystin-lr as reference) according to the method described earlier (rapala et al., 2002; rodríguez et al., 2008). the identification of microcystin/nodularin analogues and toxin amount (using microcystin-lr as reference) measurement by hplc were carried out according to the method described earlier (hautala et al., 2013). the liquid chromatography-mass spectrometry (lc-ms) and the commercial immunoassay results (quantiplate kit for microcystin, envirologix, portland, me, usa) of these samples have been published earlier, where the commercial immunoassay was referred as elisa (savela et al., 2014). from the 2009 sample panel, one additional set of aliquots (stored at -20°c) of seventeen samples were reconstituted in reagent water and analysed with the noncompetitive elisa for measuring intracellular microcystin/nodularin amount. suitable dilutions were prepared in reagent water based on the reference results to adjust the toxin concentration within the working range of the assay. results influence of temperature and incubation time on assay performance in the initial experiments the assay was performed at rt but due to relative long (2 h to overnight incubation) signal development phase, we explored the influence of the increased temperature on the assay performance. during initial bioaffinity step, temperature did not have significant effect on the assay performance; however, 1 h incubation provided ~25% higher signal as compared to 30 min incubation. based on this (1 h initial bioaffinity step at rt), we then explored the influence of the increased temperature on the rate of the signal development (fig. 1). after addition of pnpp liquid substrate, the reactions were incubated either at rt or at 37°c. ap was shown to be more fig. 1. effect of temperature and incubation time for the ap activity in non-competitive elisa using microcystin-lr as standard. the concentration (µg l–1) of microcystin-lr standard added in wells is plotted on x axis (logarithmic scale) while the corresponding absorbance at 405 nm resulting from ap activity at rt (a) or at 37°c (b) measured at different time points (30 min to 24 h) are plotted on y axis (logarithmic scale). each point represents average of two measurements. the standard errors of the means (n=2) are shown as error bars. no nco mm er cia l u se on ly 125s. akter et al. active at 37°c. within 30 min to 2 h, the incubation at 37°c yielded at least two times higher specific signal (signal blank) for microcystin-lr throughout the concentration range of 0.2-600 µg l–1 than that obtained at rt. the disadvantage of lower temperature could be compensated by prolonged incubation at rt. for example, colour development at rt with 1 h, 2-3 h and 4 h yielded similar level of signals as at 37°c using 30 min, 1 h and 2 h colour development time, respectively. unlike at rt, an increase in the background signal was observed at 37°c along prolonged incubation (fig. 1, fig. 2). performance of non-competitive elisa with different toxin variants the assay was tested for its capacity to detect different microcystins and nodularin using 11 cyanotoxin congeners in a series of concentrations (0.02 to 600 µg l–1). as indicated by the results (fig. 2), all the tested cyanotoxin analogues could be detected below the who guideline limit of 1 µg l–1 in total of 2 hours assay time (one hour sample incubation at rt followed by one hour signal development at 37°c). the detection limit (based on blank+3sd, n=56) was below 0.6 µg l–1 for all the tested toxin variants and for microcystin-lr that value was 0.2 µg l–1. from standard curves the performance range of the assay was found to be from 0.6 µg l–1 to 20 µg l–1. the signals reached plateau level with standard concentration more than 20 µg l–1and no high dose hook effect was observed within the assay with the highest standard toxin concentration of 600 µg l–1. based on the specific signal levels at 6 µg l–1, the cross reactivity for the tested microcystin/nodularin analogues relative to microcystin-lr (100%) ranged from 53% (microcystin-ly and -wr) to 107% (microcystin-lw), except for microcystin-la (30%). visual interpretation of the assay result in order to assess the performance of the assay for its applicability in near water sources for qualitative results we observed and recorded the visual colour formation at rt and at 37°c from 30 min to 24 h. in the presence of toxin, the scfv-ap becomes bound to the reaction well and then converts the colourless pnpp substrate into visually detectable yellow coloured end product (supplementary fig. 2). fig. 3 shows the comparisons of colour formation fig. 2. the non-competitive elisa standard curves for eleven different cyanobacterial toxin analogues in total 2 h assay. each point is average of duplicate and standard errors of means are not shown for visual clarity. the concentrations (0.02 to 600 µg l–1) of toxin standards added to wells are plotted in x axis (in logarithmic scale) while the corresponding signals (absorbance at 405 nm) are plotted in y axis in logarithmic scale. the detection limit (based on blank+3sd, n=56) is below 0.6 µg l–1 for all the tested toxin analogues. mc, microcystin; nod, nodularin. fig. 3. visually detectable colour formation for the non-competitive elisa. yellow visible colour was detected at rt and at 37°c in 30 min to 24 h time period using 0, 0.02-600 µg l–1 microcystin-lr standard solution. the colour intensity increases with the increase of toxin concentration at a given time point. colour development can be speeded up at higher temperature or through longer incubation. the arrows indicate the wells where 1 µg l–1 of microcystin-lr standard solution was added. no nco mm er cia l u se on ly 126 non-competitive broad specificity elisa for cyanotoxins at different times and temperatures. the who guideline value was visually detectable after 2 h colour formation at 37°c while 3-4 h incubation was required at rt. performance of non-competitive elisa with spiked water samples four spiked water samples were used in the non-competitive elisa directly (without any concentration or dilution steps) to measure the microcystin-lr concentration. the non-spiked controls were also measured. tab.1 shows the measured concentration and the recovery percentage. the reagent water and the tap water samples were found to be free of detectable toxin while the river water was found to contain a low amount of toxin (0.17 µg l–1 to 0.21 µg l–1) by all the three methods. the toxin content in the lake water was undetectable by the non-competitive elisa and the commercial elisa. however, it contained a low amount of toxin (0.03 µg l–1), as revealed by the trf assay (akter et al., 2016). the recovery values ranged from 64% to 101% and the coefficient of variation % (cv%) values of the measurements were below 10.5 for the spiking concentrations of 0.5 to 4 µg l–1. in the case of the lowest spiking concentration (0.25 µg l–1), close to the assay’s detection limit, the cv% values were below 24.5. performance of non-competitive elisa with environmental samples seventeen surface water samples (extracted intracellular toxin) were analysed by the non-competitive elisa, ppia and hplc (tab. 2). for these samples, commercial immunoassay and lc-ms results also were available (savela et al., 2014). the toxin concentration detected by the four reference methods (commercial immunoassay, ppia, hplc and lc-ms) ranged from non-detectable to as high as 40.9 µg l–1 in the samples. from these, eleven samples were found to contain less than 1 µg l–1 of toxin and two samples were found to have more than 1 µg l–1 of toxin by all the four methods. the toxin content of these samples revealed by the non-competitive elisa ranged from non-detectable to 39.1 µg l–1 and correlates well with the values obtained by the reference methods. coefficients of determination (r2) values ranged from 0.90 to 0.99 for the four reference methods. discussion immunoassays provide an easy-to-access and affordable option for quantitative detection of specific compounds. allowing reliable analysis outside well-equipped tab.1. performance of non-competitive elisa with spiked water sample. origin of water sample microcystin-lr microcystin-lr cv of the recovery and date of collection added to the sample determined by measurement (%) (µg l–1) non-competitive elisa (%) (µg l–1) 1 reagent water 0 0.25 0.22 21.4 87 0.5 0.45 3.3 90 1 0.96 2.6 96 2 1.89 2.2 94 4 3.68 1.9 92 2 drinking tap water 0 03.07.2016 0.25 0.20 24.4 80 0.5 0.40 0.0 81 1 0.88 1.8 88 2 1.87 1.2 93 4 3.59 1.5 90 3 surface water 1 (river) 0 0.21 5.6 paimio river, palikainen, 0.25 0.37 4.6 64 somero, finland 0.5 0.60 1.1 78 31.07.2009 1 1.08 3.0 87 2 2.06 0.7 93 4 3.67 10.5 86 4 surface water 2 (lake) 0* paalijärvi, riihimäki, finland 0.25 0.23 9.4 82 05.08.2009 0.5 0.49 2.9 92 1 1.04 0.5 101 2 1.96 1.5 97 4 3.84 2.2 95 *unspiked lake sample contained low amount of toxin (0.03 µg l–1) according to the trf assay method (akter et al., 2016). no nco mm er cia l u se on ly 127s. akter et al. tab. 2. intracellular microcystins/nodularins concentrations and toxin variants in environmental water samples from finland and estonia detected by non-competitive elisa and four different reference methods. place and date concentration (µg l–1) of microcystin/nodularin (intracellular) observed in microcystin-lr equivalent microcystin/nodularin non-competitive ppia hplc commercial lc-ms* variant elisa immunoassay* hplc lc-ms° lemböte byträsk, 0.40 0.17 nd 0.47 0.32 mc-yr, lemböte, åland islands, mc-dmlr finland 29.7.2009 hauninen reservoir, 0.18 0.14 0.10 0.39 0.27 mc-dmrr mc-dmrr raisio, finland 14.7.2009 hauninen reservoir, 0.65 0.26 0.13 1.14 0.86 mc-dmrr, mc-dmrr, raisio, finland mc-lr mc-rr, 15.9.2009 mc-dmlr, 1031,5 hauninen reservoir, 1.31 0.60 0.36 2.20 1.90 mc-dmrr, raisio, finland mc-dmlr, 29.9.2009 1031,5 hauninen reservoir, 0.58 0.20 0.14 1.20 0.68 mc-dmrr mc-dmrr, raisio, finland mc-dmlr, 29.10.2009 1031,5 paimio riverc, palikainen, nd 0.11 nd 0.01 nd somero, finland 31.7.2009 savojärvi, 39.13 19.40 32.50 30.40 40.90 mc-dmrr, mc-didmrr, pöytyä, finland mc-rr, mc-dmrr, 7.8.2009 mc-dmlr, mc-didmlr, mc-lr mc-dmlr maaria reservoir, 0.83 0.18 0.76 0.97 0.87 mc-rr, mc-dmrr, turku, finland mc-lr mc-rr, 11.8.2009 mc-yr, mc-lr paalijärvi,# nd nd nd 0.04 nd riihimäki, finland 5.8.2009 tuusulanjärvi, nd nd nd 0.04 nd tuusula, finland 16.9.2009 littoistenjärvi, nd 0.08 nd 0.04 0.01 mc-dmrr kaarina, finland 26.6.2009 littoistenjärvi, nd 0.20 nd nd nd kaarina, finland 04.08.2009 littoistenjärvi, 0.66 0.40 0.20 0.76 0.50 mc-rr mc-dmrr, kaarina, finland mc-rr, 3.9.2013 mc-yr, mc-dmlr, mc-lr littoistenjärvi, 5.18 9.00 3.50 7.70 3.70 mc-dmrr, mc-dmrr, kaarina, finland mc-rr, mc-rr, 11.9.2009 mc-lr mc-yr, mc-dmlr, mc-lr lake peipus, 0.73 0.24 1.10 0.55 0.60 mc-dmrr, mc-dmrr, rannapungerja beach, mc-rr, mc-rr, estonia mc-lr mc-yr, 25.8.2009 mc-dmlr, mc-lr lake peipus, 0.28 0.20 nd 0.29 0.20 mc-dmrr, mustvee beach, mc-rr, estonia mc-dmlr, 14.8.2009 mc-lr stroomi rand (sea),§ 0.37 0.17 0.17 0.34 0.25 nod-r mc-dmrr, estonia nod-r 18.8.2009 *commercial immunoassay [quantiplate kit for microcystin (envirologix)] and the lc-ms results were published earlier (savela et al., 2014);°main toxin variants are highlighted in bold; #corresponding raw water samples collected from these sources were used in spiking experiment; §commercial immunoassay, ppia, hplc and the lc-ms results for this sample was published earlier (akter et al., 2016). mc, microcystin; nod, nodularin; nd, not detected. no nco mm er cia l u se on ly 128 non-competitive broad specificity elisa for cyanotoxins high-level laboratories, immunoassays have been useful tools, for example, for environmental monitoring often performed close to the site of sampling. however, there are a number of different immunoassay configurations (wild, 2013) varying significantly e.g., in terms of the complexity of the assay procedure and instrumentation needed, and thereby also in terms of the laboratory settings required. we have recently described a straightforward time-resolved fluorometry (trf) based immunoassay for generic detection of cyanobacterial toxins, microcystins and nodularins (akter et al., 2016). in the current study a similar broad-spectrum assay for microcystins/nodularins was established in a very easily accessible and affordable elisa format facilitating the use of the assay also in resource poor settings by avoiding the need of the instrument required for trf detection. the capacity of the non-competitive elisa for generic detection of microcystins and nodularins was demonstrated using eleven commonly occurring cyanotoxin analogues (microcystin-lr, -dmlr, -rr, -dmrr, -yr, la -ly, -lf -lw, -wr, and nodularin-r). the detection limit (based on blank+3sd response) for all tested toxin analogues fell below 0.6 µg l–1 readily meeting the who guideline value of drinking water (1 µg l–1). the tested toxin analogues represent well the chemical diversity found in microcystins. for instance, both large and small as well as both polar or nonpolar amino acids can be found in the typical diversity displaying positions among the tested analogues. the fact that all these microcystin analogues, and also a penta-peptide nodularin analogue could be measured suggests that the assay is able to detect many other naturally occurring cyanotoxin analogues. we analysed the capability of the non-competitive elisa to detect the toxins in water samples using both spiked and real environmental specimens. as indicated by the acceptable recoveries obtained using raw surface water spiked with toxin (microcystin-lr) in concentrations (0.25-4 µg l–1) close to the who guideline value for drinking water, the assay can be readily applied for the analysis of both drinking and environmental surface water. the non-competitive elisa was also tested for its capacity to detect intracellular microcystin/nodularin content from 17 extracted lyophilized samples originating from different natural surface water sources. based on hplc and lc-ms, the predominant toxin variants in these samples were microcystin-rr and -dmrr. other detected toxin analogues included microcystin-lr, dmlr, -yr, -didmrr and nodularin-r. very good correlations with the reference methods, ppia, commercial immunoassay, hplc and lc-ms were observed (coefficients of determination, r2>0.90) indicating the practical applicability of the assay for samples having different microcystin analogues and nodularin. the performance of the assay is not only affected by the duration of the bacterial alkaline phosphatase catalyzed signal development step, but also the temperature during the enzymatic reaction. microcystin-lr could be detected below who guideline value using 30 min incubation at rt; however, to guarantee that all the tested toxin analogues were detected with this sensitivity 2 h incubation was required (data not shown). the detection limit could be pushed further down by extending the incubation time; after overnight incubation at rt the detection limit falls below 0.25 µg l–1 for all the tested variants. the enzyme catalyzed dephosphorylation of pnpp substrate can be significantly accelerated by increasing temperature; with 1 h incubation at 37°c all the tested variants were detectable below 0.6 µg l–1. nevertheless, prolonged colour formation at rt yields similar or even somewhat improved sensitivity due to nearly constant background signal. an additional advantage of incubating at rt is that instruments with temperature control are not needed. compared to the previously reported trf assay (akter et al., 2016), the elisa based assay shows somewhat lower sensitivity (~0.1 µg l–1 of microcystin-lr vs ~0.2 µg l–1 of microcystin-lr) and is more time consuming (10 min vs 2 h). on the other hand, the elisa test is very easy to perform and can be read with a simple elisa reader also available as portable versions, or even by naked eye if qualitative read-out is sufficient. in addition, the described elisa is economical, not only due to the inexpensive detection instrument, but also for the exceptionally affordable assay components. while the capture adda specific monoclonal antibody was obtained from commercial sources, the secondary, anti-immunocomplex, antibody was produced in a simple bacterial expression culture as a ready-made conjugate with the enzymatic label. a milligram amount of the scfv-ap protein, sufficient for thousands of assay reactions can be isolated from a 50-ml culture of e. coli by a single his-tag based affinity purification step. moreover, the immunocomplex formation based assay concept allows by-passing the production of a labeled conjugate of the analyte. this often cumbersome process is essential for competitive assays which are typically used for the detection of low-molecular-weight compounds including the cyanobacterial toxins. an additional benefit of the assay, obtained irrespective of the detection system used, is the lack of high dose hook effect. as the anti-immunocomplex binder recognize neither the free toxin nor the naked anti-adda antibody, excess of antigen does not lead to the collapse of the signal unlike in a conventional one-step (i.e., sample and tracer in the same incubation) sandwich immunoassay (davies, 2013; park and kricka, 2013). owing to this attribute, possible high dose samples cannot be misinterpreted as false negative result and only a single dilution of the sample needs to be tested during the initial screening. when quantitative results are needed, only the samples which give very high signal beyond the working range of no nco mm er cia l u se on ly 129s. akter et al. the assay need to be retested with two to three more dilutions saving time and overall cost. the non-competitive elisa concept could potentially be applied to the development of simple detection tools for various other cyanobacterial or algal toxins. the limiting factor, however, is the availability of a suitable pair of binders encompassing a primary capture (not a polyclonal) and a recombinant anti-immunocomplex antibody. if a well-performing capture antibody for a toxin exists, a recombinant antibody library can be explored e.g., by phage display to obtain the anti-immunocomplex binder (akter et al., 2016). conclusions we have here demonstrated a recombinant anti-immunocomplex antibody based non-competitive elisa for generic detection of microcystins and nodularins. the assay, validated against reference methods, is easy-to-use, robust and cost-effective, and it readily meets the who guideline level for drinking water. we believe that the assay, which can be performed with relatively simple instruments, or even qualitatively interpreted by naked eye, is wellsuited for use in water analysis laboratories, especially in the resource poor settings and at sampling location. acknowledgments this work has been financially supported (grant 823/31/2014) by the national technology agency of finland (tekes). sultana akter received personal research grants from mvtt, maaja vesitekniikan tuki (grant 24808) and from turku university foundation (grant 10146). sa, mv and ul are inventors in a pending patent application pct/fi2016/050911 concerning the anti-immunocomplex antibody described in the manuscript. the assignee of the application is the university of turku. the authors would like to acknowledge the european cooperation in science and technology, cost action es 1105 “cyanocostcyanobacterial blooms and toxins in water resources: occurrence, impacts and management” for adding value to this study through networking and knowledge sharing with european experts and researchers in the field. we are grateful to those colleagues who provided samples: sonja nybom (åbo akademi university), pirkko ala-uotila (raisio-naantali waterworks), pirkko pajakko (turku municipal water company), marko järvinen (finnish environment institute), kirsti lahti (water protection association of the river vantaa and helsinki region) and aune annus and coworkers (tervisekaitse, estonia). references akter s, vehniäinen m, spoof l, nybom s, meriluoto j, lamminmäki u, 2016. broad-spectrum noncompetitive immunocomplex immunoassay for cyanobacterial ceptide hepatotoxins (microcystins and nodularins). anal. chem. 88:10080-10087. botes dp, tuinman aa, wessels pl, viljoen cc, kruger h, williams dh, santikarn s, smith rj, hammond sj, 1984. the structure of cyanoginosin-la, a cyclic heptapeptide toxin from the cyanobacterium microcystis-aeruginosa. j. chem. soc. perkin trans. 1:2311-2318. choi bw, namikoshi m, sun f, rinehart kl, carmichael ww, kaup am, evans wr, beasley vr, 1993. isolation of linear peptides related to the hepatotoxins nodularin and microcystins. tetrahedron lett. 34:7881-7884. codd ga, morrison lf, metcalf js, 2005. cyanobacterial toxins: risk management for health protection. toxicol. appl. pharmacol. 203:264-272. dahlem am, 1989. structure/toxicity relationships and fate of low molecular weight peptide toxins from cyanobacteria. phd thesis, university of illinois, urbana-champaign. davies c, 2013. principles of competitive and immunometric assays (including elisa), p. 29-59. in: d. wild (ed.), the immunoassay handbook, 4th ed. elsevier. deshpande ss, 1996. enzyme immunoassays: from concept to product development. springer, new york: 464 pp. grosse y, baan r, straif k, secretan b, el ghissassi f, cogliano v, 2006. carcinogenicity of nitrate, nitrite, and cyanobacterial peptide toxins. lancet oncol. 7:628-629. hautala h, lamminmaki u, spoof l, nybom s, meriluoto j, vehniainen m, 2013. quantitative pcr detection and improved sample preparation of microcystin-producing anabaena, microcystis and planktothrix. ecotoxicol. environ. saf. 87:49-56. huovinen t, syrjanpaa m, sanmark h, brockmann ec, azhayev a, wang q, vehniainen m, lamminmaki u, 2013. two scfv antibody libraries derived from identical vl-vh framework with different binding site designs display distinct binding profiles. protein eng. des. sel. 26:683-693. lawton la, edwards c, 2008. conventional laboratory methods for cyanotoxins. adv. exp. med. biol. 619:513-537. mazur-marzec h, meriluoto j, plinski m, szafranek j, 2006. characterization of nodularin variants in nodularia spumigena from the baltic sea using liquid chromatography/mass spectrometry/mass spectrometry. rapid commun. mass spectrom. 20:2023-2032. merel s, walker d, chicana r, snyder s, baures e, thomas o, 2013. state of knowledge and concerns on cyanobacterial blooms and cyanotoxins. environ. int. 59:303-327. meriluoto j, codd ga, 2005. toxic: cyanobacterial monitoring and cyanotoxin analysis. åbo akademi university press, åbo: 149 pp. neilan ba, dittmann e, rouhiainen l, bass ra, schaub v, sivonen k, borner t, 1999. nonribosomal peptide synthesis and toxigenicity of cyanobacteria. j. bacteriol. 181:4089-4097. niedermeyer t, 2014. microcystin congeners described in the literature. figshare. http://dx.doi.org/10.6084/m9.figshare. 880756 nishiwaki-matsushima r, ohta t, nishiwaki s, suganuma m, kono nco mm er cia l u se on ly 130 non-competitive broad specificity elisa for cyanotoxins hyama k, ishikawa t, carmichael ww, fujiki h, 1992. liver tumor promotion by the cyanobacterial cyclic peptide toxin microcystin-lr. j. cancer res. clin. oncol. 118:420-424. park jy, kricka lj, 2013. interferences in immunoassay, p. 403416. in: d. wild (ed.), the immunoassay handbook, 4th ed. elsevier. puddick j, 2013. spectroscopic investigations of oligopeptides from aquatic cyanobacteria: characterisation of new oligopeptides, development of microcystin quantification tools and investigations into microcystin production. phd thesis, university of waikato, new zealand. rapala j, erkomaa k, kukkonen j, sivonen k, lahti k, 2002. detection of microcystins with protein phosphatase inhibition assay, high-performance liquid chromatography-uv detection and enzyme-linked immunosorbent assay: comparison of methods. anal. chim. acta 466:213-231. rinehart k, namikoshi m, choi b, 1994. structure and biosynthesis of toxins from blue-green algae (cyanobacteria). j appl. phycol. 6:159-176. rinehart kl, harada k, namikoshi m, chen c, harvis ca, munro mhg, blunt jw, mulligan pe, beasley vr, dahlem am, carmichael ww, 1988. nodularin, microcystin, and the configuration of adda. j. am. chem. soc. 110:8557-8558. rodríguez em, acero jl, spoof l, meriluoto j, 2008. oxidation of mc-lr and -rr with chlorine and potassium permanganate: toxicity of the reaction products. water res. 42:17441752. savela h, vehniäinen m, spoof l, nybom s, meriluoto j, lamminmäki u, 2014. rapid quantification of mcyb copy numbers on dry chemistry pcr chips and predictability of microcystin concentrations in freshwater environments. harmful algae 39:280-286. self ch, thompson s, street t, lamb kj, duffin g, dessi jl, turnbull m, 2013. non-competitive immunoassays for small molecules-the anti-complex and selective antibody systems, p. 61-65. in: d. wild (ed.), the immunoassay handbook, 4th ed. elsevier. sivonen k, 1996. cyanobacterial toxins and toxin production. phycologia 35:12-24. sivonen k, 2008. emerging high throughput analyses of cyanobacterial toxins and toxic cyanobacteria. adv. exp. med. biol. 619:539-557. sivonen k, jones g, 1999. cyanobacterial toxins, p. 41-111. in: i. chorus and j. bartram (eds.), toxic cyanobacteria in water: a guide to their public health consequences, monitoring and management. taylor & francis. spoof l, catherine a, 2017. appendix 3. tables of microcystins and nodularins, p. 526-537. in: j. meriluoto, l. spoof, g. codd (eds.), handbook of cyanobacterial monitoring and cyanotoxin analysis. j. wiley & sons. stewart i, seawright aa, shaw gr, 2008. cyanobacterial poisoning in livestock, wild mammals and birds–an overview. adv. exp. med. biol. 619:613-637. stotts rr, namikoshi m, haschek wm, rinehart kl, carmichael ww, dahlem am, beasley vr, 1993. structural modifications imparting reduced toxicity in microcystins from microcystis spp. toxicon 31:783-789. sueoka e, sueoka n, okabe s, kozu t, komori a, ohta t, suganuma m, kim sj, lim ik, fujiki h, 1997. expression of the tumor necrosis factorα gene and early response genes by nodularin, a liver tumor promoter, in primary cultured rat hepatocytes. j. cancer res. clin. oncol. 123:413-419. who, 2011. guidelines for drinking water quality, 4th ed. who, geneva: 564 pp. wild d, 2013. immunoassay for beginners, p. 7-10. in: d. wild (ed.), the immunoassay handbook, 4th ed. elsevier. no nco mm er cia l u se on ly layout 1 introduction mangrove forests are located in intertidal and transitional zones between aquatic and terrestrial environments (lee et al., 2006) with special ecological conditions (thu and populus, 2007) and occur in up to 70% of tropical and subtropical coastlines worldwide. they are important for their ecological values, highly production (ghosh et al., 2010), economic activities such as fisheries and aquaculture and biogeochemical cycles (thu and populus, 2007; zahed et al., 2010). they provide breeding and feeding areas for many macroand micro-organisms. the microbial community living in the sediments of mangrove forests are exposed to several spatially and temporally variable ecological, biogeographical and anthropogenic fetors, including abundance of organic and inorganic matter, food web structure, nutrient cycling and pollution (ghosh et al., 2010). since mangrove ecosystems endure periodic tidal flooding, environmental factors, such as nutrient availability and salinity, are highly variable, which make mangroves a unique ecosystem with specific characteristics (holguin et al., 2006). qeshm island is the largest island in the persian gulf (salehipour milani and jafar beglu, 2012), and its mangrove forests, covering about 67.5 km² and located within the northwest estuaries of the island, in khoor-ekhooran and sandy islands apposite villages of tabl and laft to gooran (55° 63’e 26° 75’n and 55° 80’e 26° 93’n), are protected areas, international wetlands and biosphere reserves (zahed et al., 2010). they are also the largest area of mangrove forests in the northern part of the persian gulf (khosravi, 1992). the persian gulf is experiencing stressful conditions due to high seasonal variations in temperature (<10°c in winter and >30°c in summer) (khosravi, 1992; zehzad and majnounian, 1997). annual rainfall in iranian mangrove forests is typically <200 mm, evaporation exceeds rainfall (joekar and razmjoo, 1995) and, thus, salinity is high (38 to 50) (zahed et al., 2010; salehipour milani and jafar beglu, 2012). as a consequence, life in these environments shows a high tolerance to salinity and heat stress (yan and sinica, 2007). the hypersaline environments which are derived from the sea could be divided into four groups according to the salinity range and organism origin: i) 6-7 to 10% salinity includes originally marine various biota, ii) 10-14% salinity includes usually halotolerant, halophilic or adapted biota, iii) 14-30% salinity includes both extremely and moderately halophilic bacteria, iv) more than 30% salinity contains only bacterial community living on the organic matter (por, 1980). according to the microorganism’s ranges of salinity tolerance, they are categorized into three groups: slightly, moderately and extremely halophiles (ventosa et al., 2006; zhang et al., 2014). halophilic bacteria can be present in almost all hypersaline environments including marine or terrestrial, anaerobic or aerobic, polar and tropical, or acidic to alkaline environments (javor, 2012). microorganisms from extreme environments are able to enhance oil recovery and bioremediation (lima et al., 2011), and produce important secondary compounds with potential to be used in pharmaceutical industries, article isolation and identification of halophilic and halotolerant bacteria from the sediments of the qeshm island mangrove forest pegah javid,1 hassan zadabbas shahabadi,2 homeyra amirkhani,3 narges amrollahi biuki,1,4* mohammad sharif ranjbar1 1department of marine biology, faculty of marine science and technology, university of hormozgan, bandar ‘abbas; 2department of marine biology, faculty of marine sciences, chabahar maritime university, chabahar; 3microbiology department; food, drug and standard bureau of qeshm free zone, qeshm island; 4department of modern technology, mangrove forest research center, university of hormozgan, bandar ‘abbas, iran abstract due to specific environmental and ecological conditions, mangrove forests are known as marine transitional zones between sea and land, and, as such, they host organisms with high ecological plasticity. the mangrove forests of qeshm island (iran) are relatively pristine habitats and represent an ideal target for investigating patterns of either aquatic or benthic biodiversity. to provide insights on microbial diversity in this area, nineteen halophilic and halotolerant bacteria were isolated from the sediments in 2017 during low tide. the extracted bacterial strains were studied morphologically by streaking, initial observation of colonies and bacterial staining, and characterized using a battery of biochemical tests including koh, mr, vp, urease, tsi, s/i/m, mac, lia, odc, adh, oxidase, catalase, and tryptophan deaminase. the optimum growth of halophilic bacteria was observed in salt concentrations from 5 to 20% nacl, whereas the extreme halophilic gram-positive strain grew in salt concentration of up to 30% nacl. molecular analyses were also carried out on four halophilic strains and one extreme halophilic gram-positive bacteria. phylogenetic taxonomy analysis, after 16s rdna gene sanger sequencing, revealed that the halophilic bacteria were closely related to the strain types of the genus bacillus including bacillus licheniformis, bacillus velezensis, bacillus paralicheniformis and bacillus sp. with 99% bootstrap value. the extreme halophilic strain was associated to strains of planococcus plakortidis with 100% bootstrap value. no nco mm er cia l u se on ly p. javid et al.2 agriculture, etc. (janek et al., 2010). isolated strains including bacillus sp. from more tough environments, such as marine and salty sediments can produce surfactin products (kiran et al., 2010; donio et al., 2013). as an example, bacillus subtilis is able to produce a cyclic lipopeptide surfactin, which is known as one of the most powerful surfactants (kiran et al., 2010). this product is active at extreme salinity, temperatures and ph levels (pacwa-plociniczak et al., 2011). other biological products, such as biosurfactants and antibiotics, are also produced by microorganisms living in extreme environments (abdelmawgoud et al., 2008). although the molecular evolution of stability and life of macromolecules has not been fully revealed yet (vasavada et al., 2006), the survival and proliferation of microorganisms in extreme environmental conditions could depend on their ability to produce biologically active compounds (valentine 2007), which in turn, could reasonably have a pharmaceutical importance. as an example, a pharmacologically important biosurfactant isolated from a halophilic bacillus sp. bs3 at extreme environment of solar salt work has anticancer activity in breast cancer and antiviral activity in shrimps’ white spot syndrome (donio et al., 2013). also, halophilic bacteria are suitable choices for bioremediation of hypersaline habitats (zhang et al., 2014). many studies have indicated that due to the high diversity of bacterial species, evaluation of bacterial diversity using culture-dependent methods may lead to inaccurate information. therefore, over the past two decades, many researchers have turned from culturedependent approaches to molecular (16s rrna-dependent) approaches. although this change of method has led to the discovery of many new species of microbes, bacterial cultivation, wherever possible and with exclusion of viable but non-culturable (vbnc) bacteria (li et al., 2014), remains the best method to investigate their physiology under variable ecological conditions (yeon et al., 2005). culture-dependent methods can be technologically enhanced by molecular techniques and molecular analysis. also, pcr-based methods following dna technologies could be complementary for cultivationdependent studies in order to discover the biodiversity or genetic differences of bacterial communities (yoon et al., 2003; fry, 2004; yeon et al., 2005). studies on bacteria communities of soils and sediments have been conducted worldwide. in mangrove sediments, among the others, studies have been carried out on the diversity and biotechnology potential of bacteria in brazil (dias et al., 2009), isolation of bacillus swezeyi sp. nov. and bacillus haynesii sp. nov. from negev desert of isreal (dunlap et al., 2017), bacterial biodiversity in sundarban, india (ghosh et al., 2010), recovery of novel bacteria diversity in china (liang et al., 2007), and so on. to provide insights on microbial diversity in mangrove sediments, we have isolated and characterized, for the first time ever, some bacterial strains from the sediments of the qeshm mangrove forests in iran. methods sampling sediment sampling was carried out with sterile pestles in september 2017 from the northwestern part of the mangrove forests in qeshm island, at soheili jetty (26° 78′ n, 55° 76′ e) (fig. 1) during the maximum low tide from three zones of upper intertidal near tree coverage (a); intertidal (b); and under tidal zones near the sea water (c). the sediment samples were collected in sterile glass jars, put into ice powder and then transported to the laboratory in 4°c refrigerator for further analyses. bacteria cultivation, purification, and staining sediment samples from the three zones (a, b and c) were initially dissolved in 0.85% nacl solutions. different dilutions of 10–1, 10–2 and 10–3 were then prepared and cultured in nutrient agar (na) at 30°c for 24 and 48 hours. then, the grown colonies were isolated on new na cultures for subsequent biochemical tests, staining, and molecular studies. nacl tolerance of strains and their growth were tested. more specifically, for assays of salt tolerance, nutrient broth (nb) medium was supplemented with different sodium chloride concentrations of 5%, 10%, 15%, 20%, 25% and 30% (w/v). the isolated strains in different nacl concentrations were incubated at 30°c and monitored for growth after 24 and 48 hours. the gram reaction of bacteria was determined by gram staining set (qatran shimi, iran). gram staining is a useful method for detecting the structure of bacterial cell wall. according to the wall structure of bacterial cell, after staining, they show a purple appearance in gram-positive bacteria or red in gram negative ones. staining was followed in accordance with manufacturer’s protocol. growth and common biochemical tests for identification of microorganisms were performed as follows: koh test (potassium hydroxide test) with 3% koh solution was carried out for confirming the gram staining process. catalase test was conducted for finding out whether the strains had catalase enzyme for decomposition of h2o2; mr (methyl red)/vp (voguesproskauer) test was used to distinguish and identify the general forms of bacteria and identify the final product of bacterial fermentation pathway. urease test was done in urease broth medium culture to determine which strain produced the urease enzyme that hydrolyzes urea to ammonia and carbon dioxide. sugar fermentation test was used in differential tsi or triple sugar iron agar culture medium containing three types of sugar (lactose, sucrose no nco mm er cia l u se on ly phylogeny of bacterial community of mangrove forest sediments 3 and glucose). this test was carried out to differentiate the microorganisms according to their ability in reduction of sulfur and fermentation of carbohydrates. sulfur reduction, indole production ability, and motility of strains were tested through culturing in sim (sulfur, indole, and motility) medium to differentiate the strains. culturing in macconkey agar was carried out for detection of gram-negative bacteria. the decarboxylation tests show the microorganism’s ability in removing the carboxyl group from an amino acid, and are useful for the differentiation of enterobacteriaceae. the decarboxylation test for lysine amino acid was carried out in lysine iron agar (lia). ornithine decarboxylation test was conducted in ornithine decarboxylase agar (odc). arginine dihydrolase test was done with culturing the bacteria in arginine dihydrolase broth (adh) medium to detect the bacteria producing arginine dihydrolase enzyme that decarboxylates arginine. oxidase test was carried out to identify the organisms producing cytochrome oxidase enzyme. indole test was carried out by culturing strains in tryptophan broth. this test shows the ability of bacteria in converting tryptophan into indole. dna extraction, pcr conditions, and molecular identification the bacteria were cultured in 10 cc tryptophan broth and yeast extract with 10% salinity for proliferation. dna was extracted by genomic dna extraction kit from gram-positive bacteria (gene transfer pioneers, iran) and suspended in 30 ml elution buffer. extracted dna quality was evaluated by 1.5% agarose gel electrophoresis using tris-borate-edta as the buffer and redsafetm (intron, south korea) for gel staining. the extracted dna was used as a template for amplification and sequencing. the universal bacterial primers, 27f (forward: 5’-aga gtt tga tcm tgg ctc ag-3’) (lane, 1991) and 1525r (reverse: aag gag gtg wtc car cc) (lane, 1991) were used for amplification of 16s rdna genes by polymerase chain reaction. pcr reactions were carried out in 25 μl of master mix, with 2.5 μl of 10x pcr buffer (bioron, germany), 1.25 μl of 100 mm mgcl2 (bioron, germany), 1 μl of 25 mm dntps (geneall, korea), 1 μl of each 10 pm forward and reverse primer (pishgam, iran), 5 u taq dna polymerase, and about 15 ng μl–1 of template. the 16s rdna was amplified using a dna thermal fig. 1. map of sampling site on mangrove forest, soheili jetty, the northwestern part of qeshm island, iran. credit: eghbal zobeiri. no nco mm er cia l u se on ly p. javid et al.4 cycler with the following profile: 95°c for 4 min; 38 cycles of 94°c for 1 min, 56.5°c for 1 min, and 72°c for 2 min and 30 s; and a final cycle of 72°c for 10 min. pcr products were visualized through 1% agarose gel, and the pcr products were sequenced through sanger dna sequencing method by bioneer company, south korea. the sequences were blasted in the national center of biotechnology information (ncbi) in order to identify the strains. according to the blast results in ncbi, the strains with the most similarity and high percentage of identity were selected for finding the phylogenetic taxonomy relationships. the sequencing data were aligned and edited using clustalx (larkin et al., 2007) to obtain overlapping and equal segments. then the phylogenetic analysis was done in mega 7 (kumar et al., 2016) for designing the tree and finding the relations between strains. bootstrap consensus tree was obtained by nj analysis from 1000 replicates. bootstrap analysis was done to determine the limits of the branching. the phylogenetic tree was rooted in alicyclobacillus acidocaldarius as the outgroup. results biochemical and growth features of bacterial strains from qeshm mangrove sediments, 19 cultivable strains were isolated on the basis of morphology, color, gram staining, biochemical tests, growth activity and salinity resistance. molecular identification was carried out on five out of the 19 strains. all bacteria were grampositive. tab. 1 shows the biochemical and growth features of five molecularly identified strains and tab. 2 shows the growth status of strains in different salinities. strains 1 and 2 were isolated from sediments of zone a, strain 3 was isolated from sediments of zone b, and strains 4 and 5 belonged to sampled sediments from zone c. molecular identification and phylogenetic analyses the 16s rdna gene sequences consisted of 967 bp for strain 1, 1436 bp for strain 2, 1442 bp for strain 3, 1441 bp for strain 4 and 1047 bp for strain 5. they were blasted in the ncbi database, and were compared with other species. tab. 1. biochemical and growth features of 5 molecularly identified strains. features/strains strain 1: pg-1-ms strain 2: pg-2-ms strain 3: pg-3-ms strain 4: pg-4-ms strain 5: pg-5-ms gram + + + + + koh mr vp + urease + + tsi alk/alk alk/alk alk/ac alk/ac alk/alk s/i/m / / + / / + / / + / / + / / + mac lia + + + + odc adh oxidase + + + + catalase + + ttryptophan deaminase alk, alkaline reaction; ac, acid production. tab. 2. bacterial growth in different concentrations of sodium chloride. the bold columns show the strains identified by molecular study (1: pg-1-ms; 4: pg-2-ms; 7: pg-3-ms; 16: pg-4-ms; 19: pg-5-ms). positive items show the strains’ growth in different concentrations of nacl. salinity%/strains 1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 0 + + + + + + + + + + + + + + + + + + + 5 + + + + + + + + + + + + + + + + + + 10 + + + + + + + + + + + + + + + + + + 15 + + + + + + + + 20 + + + + 25 + 30 + no nco mm er cia l u se on ly phylogeny of bacterial community of mangrove forest sediments 5 molecular identification of the 16s rdna gene sequences includes 12 strains of bacillus sp. bacillus licheniformis and bacillus paralicheniformis in clade 1 (cl1), 6 strains of bacillus velezensis in clade 2 (cl2) and 9 strains of planococcus plakortidis and planococcus maritimus in clade 3 (cl3) along with one outgroup. the analysis in this study showed that strains in the same clade are sister groups and clade 1 and 2 are supported as monophyletic groups with a bootstrap value of 100%. the blast results of cl1 showed that pg-1-ms is associated with the species of bacillus licheniformis strains of this clade with 99.90% (kj526876.1, kj526860.1, and ef472268.1) and 99.79% (mg980062.1) identities. in the same clade, the strain pg-3-ms has high identity with other strains of bacillus paralichenifirmis (100% identity with strain mh538129.1 and 99.93% identity with strains cp023666.1 and mf321832.1). meanwhile, mh538129.1 and cp023666.1 have lower identity (99.79%) with bacillus licheniformis. in addition, pg-4-ms was identified as bacillus sp., and had 100% identity with other bacillus sp. strains (mg814028.1 and mg814024.1) in the blast result, while the identity of pg-4-ms with bacillus paralicheniformis is 99.93%. another bootstrap value of 100% shows that clade 3 is paraphyletic for the previous monophyletic groups. the clade 1 contains strains 1, 3 and 4 of the recorded mangrove sediment bacteria. clade 2 has all bacillus velezensis strains including our strain 2. clade 3, as a paraphyletic group in the phylogenetic tree, includes the species planococcus plakortidis, which corresponds to our strain 5. strains pg-1-ms, pg-2-ms, pg-3-ms, and pg-4-ms belong to genus bacillus, whereas pg-5-ms belongs to genus planococcus. about 967 bp for pg-1-ms (accession number: mh847742.1) and 1442 bp for pg-3ms (accession number: mh842119.1) were analyzed in the final alignment. pg-1-ms and pg-3-ms were closely related to bacillus licheniformis and bacillus paralicheniformis, respectively (99% bootstrap value). the pg-4-ms (accession number: mh842120.1) was related to strains pg-1-ms and pg-3-ms and genus bacillus with 99% bootstrap value, but other tests and biochemical analyses showed different results from other identified species mentioned in tab. 1. the neighbor-joining phylogenetic tree (fig. 2) revealed that strain pg-2-ms (accession number: mh838019.1) belongs to species bacillus velezensis and is related to the other strains of species bacillus velezensis with 99% bootstrap value. blast results also showed that pg-2-ms has 100% identity with other strains of bacillus velezensis (mf662480.1, mf662477.1, mf662476.1, mf662472.1, and mf662470.1) presented in clade 2. the last strain pg-5-ms (accession number: mh842121.1) is associated with other species of planococcus plakortidis with 100% bootstrap value in phylogenetic taxonomy analysis. blast results revealed that pg-5-ms has more than 95.52% (mg705879.1, and mg705828.1) and 99.60% (mh384429.1, cp016539.2, nr_109414.1) identity with identity with planococcus plakortidis. although these blast results showed close relation between pg-5-ms and planococcus maritinus (mg705829.1) with 99.52% identity, biochemical characteristics showed more similarity of pg-5-ms to p. plakortidis. thus, the highest bootstrap values among our strains and other species of genus bacillus and planococcus were 99% (bacillus sp., bacillus licheniformis, bacillus paralicheniformis, bacillus velezensis) and 100% (planococcus plakortidis), respectively. strains pg-1-ms, pg-2-ms, pg-3-ms, and pg-4-ms showed common features with genus bacillus as they were all gram-positive, rod-shaped, endospore-forming, motile and aerobic. pg-1-ms and pg-2-ms grew in up to 10% salinity in nb, but not in higher nacl concentrations. pg-3-ms and pg-4-ms tolerated more salinity and grew in enriched nb with 20% nacl. the growth condition in nb enriched by nacl, interestingly, showed that pg-5-ms is an extreme halophilic strain that grew in 30% nacl concentration. strain pg-1-ms belongs to aerobic, endospore-forming motile rod bacteria that produced yellow glossy round shaped with regular margin colonies in the na. this halotolerant strain was molecularly identified as a strain of bacillus licheniformis (more than 99.79% identity with other strains of bacillus licheniformis in clade 1; accession number: mh837742.1). strain pg-2-ms, molecularly identified as a strain of bacillus velezensis (100% identity with other strains of bacillus velezensis in clade 2 and 100% bootstrap value; accession number: mh838019.1), was aerobic diplococcus bacteria with creamy white flat, viscose and irregular-margin colonies in the na. strain pg3-ms, recorded as a strain of bacillus paralicheniformis (more than 99.93% identity with other strains of bacillus paralicheniformis in clade 1; accession number: mh842119.1), was an aerobic bacteria that formed convex, multilayered mounds of flower-like or amorphous, slime, approximately colorless (very light creamy-white) colonies with 2-3 mm in diameter in the na culture. strain pg-4ms, identified as a strain of bacillus sp. (99.93% identity with other strains of bacillus sp. in clade 1; accession number: mh842120.1), formed spreading edged mucoid transparent light pinkish colonies on na culture. strain pg5-ms, the strain of planococcus plakortidis (more than 99.5% blast identity and 100% bootstrap value; accession number: mh842121.1), was gram-positive, non-spore forming, slow-growing, aerobic coccoid bacteria, and produced colonies with approximately 1 mm in diameter after about 72 hours at 30°c. the colonies were smooth, glossy, circular, small and round-shaped with regularmargin in orange on na culture. no nco mm er cia l u se on ly p. javid et al.6 discussion and conclusions mean seawater salinity is ca. 35, and many microorganisms which live at this salinity can endure slightly higher salinities, whereas only halotolerant and halophilic species can grow in extremely hypersaline environments (javor, 2012). although, in comparison to other habitats, hypersaline environments possess low prokaryotic diversity (guixa-boixareu et al., 1996), they are home to a high diversity of halotolerant bacteria and extremely halophilic bacteria (yoon et al., 2003). the microorganisms that tolerate very high salinity values show different metabolic features adapted to such extreme conditions (zhang et al., 2014). the genus bacillus consists of various species with some general characteristics: rod-shaped, and chainforming gram-positive bacilli which can grow anaerobically or aerobically (thwaite and atkins, 2012), the majority of them being motile and having lateral flagella (stoica and ionut, 2017). their function in biochemical tests, growth features and even the phenotype of sporangia can influence their classification, and they can cope with unfavorable growth conditions due to their endospores (zeigler and perkins, 2015). members of the genus bacillus possess variable features and can live in different environments, or endure environmental variations of oxygen availability, humidity, temperature and nutrient accessibility (nicholson, 2002). for example, different species of bacillus might be aerobic or anaerobic, grampositive or gram-variable and live in all environments from dust, soil and water (la jeon et al. 2012) to skin and human gut (mandell et al., 2010). bacillus anthracis and bacillus cereus are human pathogens (farrar, 1963). bacillus cereus populations are endosymbionts in the gut of cockroach species blaberus giganteus and have nutritional advantage for these insects (feinberg et al., 1999). it is reported that the species bacillus infernus has been found nearly 3 km below the surface of earth (boon et al., 1995). some even fig. 2. neighbor-joining phylogenetic tree for 27 strains, including qeshm mangrove sediment species (pg-1-ms to pg-5-ms), based on 16s rdna sequences. the numbers beside the branches are bootstrap values with 1000 replications. alicyclobacillus acidocaldarius is selected as the outgroup. no nco mm er cia l u se on ly phylogeny of bacterial community of mangrove forest sediments 7 could be found in extreme salty ecosystems such as solar salt works and condenser water with 155 ppt of salinity (donio et al., 2013). any species might include different strains, which are sympatric (palmisano et al., 2001), so it is obvious that different strains of a species do not show the same features. although there are strains of bacteria which are not potentially halophilic, with the passage of time, they can tolerate salinity as they adapt to their environment (zhang et al., 2014). the species of genus bacillus have been increasingly important in applied microbiology and biochemical studies such as production of natural products, antibiotics, surfactants and enzymes (ruiz-garcia et al., 2005; jasim et al., 2016). they have been studied also as probiotics in fish food and used as feeding supplement in aquaculture (martínez cruz, 2012; ramesh et al., 2015; meidong et al., 2017) and livestock (kaewtapee et al., 2017) for improving performance and health in animals (meng et al., 2010). especially, bacillus licheniformis and bacillus subtilis have been used as pig supplementary diet (larsen et al., 2014; jørgensen et al., 2016). different species of genus bacillus could be found in various habitats from nature and plants (arnesen et al., 2008; jasim et al., 2016), and in intestinal tract of insects and mammals (arnesen et al., 2008) to soil (palmisano et al., 2001; dunlap et al., 2017) and water (arjmand et al., 2016). bacillus enfolds about 70 species (thwaite and atkins, 2012) four of which have been recently presented as surfactant-producing bacteria, including bacillus licheniformis (weigmann, 1898) and bacillus velezensis (ruiz-garcia et al., 2005) isolated from velez river in southern spain for the first time, bacillus paralicheniformis (dunlap et al., 2015) that have been isolated from fermented soybean paste for the first time, and bacillus sp. which have been studied in the present paper. as it was mentioned before, according to the continuously increasing number of identified bacteria with high similarity to each other, recognition via classical identification methods such as analysis of physiological and morphological characteristics, the composition of cell walls and biochemical features is very difficult (fan et al., 2017). there might be some other features other than morphology and biochemistry characteristics which make microorganisms distinguished from each other such as dna sequences, fatty acid composition and even resistance to genetic transformation; these kinds of differences might exist even among the species of the same genus (bacon and hinton, 2011). thus, there might exist some biochemical features which vary among species of a genus or strains of a species. the genus planococcus was identified more than one century ago by migula (1894), and was later modified by nakagawa et al. (1996) (migula, 1894; nakagawa et al., 1996; kaur et al., 2012). some species of this genus are alkaliphilic and thermotolerant. a novel strain of species planococcus plakortidis was isolated from a marine sponge, plakortis simplex (kaur et al., 2012) and received the authority by kaur et al. (2012). planococcus plakortidis isolated from a genus of marine sponge showed positive activity for both catalase and oxidase (kaur et al., 2012) while planococcus sp. mc01 (99.3% nucleotide identity with planococcus plakortidis) (ma et al., 2013) and planococcus maritimus (yoon et al., 2003) were found to be catalase-positive and oxidase-negative. additionally, in our study, planococcus plakortidis pg-5-ms had negative and positive activity for catalase and oxidase, respectively. our phylogenetic taxonomy analysis using 16s rdna through nj phylogenetic tree showed that clade cl1 includes bacillus licheniformis, bacillus paralicheniformis, bacillus sp. and their similar strains in a species, clade cl2 consisted of bacillus velezensis and similar strains and clade cl3 possess planococcus plakortidis and its strains (fig. 2). the affinity of molecularly identified strains in trees showed 99% or more bootstrap value with other strains. bacillus licheniformis is a halophilic species, and different studies have shown that the presence of nacl in the medium has better effects on bacterial growth and the production of bacterial products. there was positive growth on the medium enriched by 7% nacl (v/w), and the grown bacteria successfully produced tannin acyl hydrolase (mondal and pati, 2000). nacl concentration in the modified medium of a strain of bacillus licheniformis isolated from saline soil had positive effect on γ-pga production resulting in the most volumetric yield of this product (13.86 g l–1) as the medium was enriched by 8% nacl; the strain had positive growth in 12% concentration of nacl as well (wei et al., 2010). interestingly, the phylogenetic analysis has also revealed that two strains of genus bacillus (bacillus swezeyi and bacillus haynesii) had a close relationship with the clade that included bacillus licheniformis and its members, and they showed tolerance to up 12% nacl (w/v) (dunlap et al., 2017). shivaji et al. (2006) showed that bacillus licheniformis forms white irregular, fried-egg-like colonies on na, tolerated up to 11.6% nacl (v/w) but not 17.4% and 23.4% concentrations of nacl (v/w). bacillus velezensis showed growth capability at 12% w/v nacl concentration and a temperature range of 1545 °c; also, grew on tsa medium while producing rough white-creamy colonies with irregular edges (ruiz-garcia et al., 2005), which was not far from our observations on this strain grown in na. bacillus velezensis was reported as a bacterium which acts as a plant growth promoting strain (fan et al., 2017). presence of strain bacillus velezensis in the collected sediments close to the plants in our research further proves the role of this strain in rhizosphere colonization. dunlap et al. (2015) recorded a strain of bacillus paralicheniformis as absolute anaerobic no nco mm er cia l u se on ly p. javid et al.8 gram-positive, motile strain that can form creamy mucoid, semitransparent colonies with 3-4 mm in diameter on r2a agar that can tolerate nacl concentrations up to 10% (w/v) (dunlap et al., 2015). planococcus plakortidis was identified as a grampositive coccoid strain for the first time in 2012 from the bay of bengal (kaur et al. 2012). planococcus sp. has been reported as a halotolerant and alkaliphilic slowgrowing bacterium with orange-red colonies and 2-3 mm in diameter that grew after two days while the optimal temperature was 30 °c; in this study, the growth occurred at from 2% to 17% (w/v) nacl (ma et al., 2013). the strains of planococcus maritimus with 100% bootstrap value in cl3 with planococcus plakortidis and its strains in our study showed high affinity. planococcus maritimus was isolated from seawater of the tidal zone in korea for the first time. the biochemical characteristics of this species, however, showed either similarities (grampositive, cocci, motile, positive growth in 0% nacl, negative urease) or dissimilarities (positive catalase, negative oxidase, tolerant to different salinities) (yoon et al., 2003) to the putative planococcus plakortidis strain isolated in our study. positive growth in 15-17% (v/w) concentration of nacl by planococcus maritimus in yoon et al. (2003) survey and the same result in up to 12% nacl (v/w) by the strains of planococcus sp. in dunlap et al. (2017) study showed that this genus is halotolerant. thus, based on our findings, we conclude that the identified halophilic bacteria isolated from the sediments of the mangrove forests in qeshm island, south of iran, were mostly from genus bacillus and planococcus, which possess variant species and strains with different degrees of tolerance to salinity. our phylogenetic analysis showed a close relationship between different identified strains with previous studies and recorded strains. strains pg-1-ms, pg-3-ms and, pg-4-ms in cl1, strain pg-2-ms in cl2 and strain pg-5-ms in cl3 lie in monophyletic groups. the noted strains generally formed clades in a phylogenetic tree with more than 98% bootstrap value protection. the results presented in our spot survey in the mangrove sediments of qeshm island, south of iran, confirm that this ecosystem hosts a peculiar diversity of halophilic bacteria with bio-ecological and functional characteristics either similar or different from those of similar strains isolated from other environments. the results presented here suggest that the prokaryotic diversity of this peculiar ecosystem deserves further attention, also in order to identify microorganisms that could possess a valuable potential in bioremediation and better growth of halophile plants. acknowledgments this research was granted by the university of hormozgan. we are grateful to kazem ghorbani, the head of drug and standard bureau of qeshm free zone, for his kind permission for working in the microbiology department, and to somayeh kouchaki for her efforts in coordination. we also thank somayeh mohammadi for her valuable contribution at the microbiology department. we are grateful to eghbal zobeiri for designing the map picture. the authors would also like to thank novin biotechnology department of isfahan for provision of facilities during molecular experiments. corresponding author: amrollahi@hormozgan.ac.ir key words: halophiles; extreme halophile; halotalerant; grampositive; molecular identification; mangrove sediments. received: 8 december 2019. accepted: 16 march 2020. this work is licensed under a creative commons attribution noncommercial 4.0 license (cc by-nc 4.0). ©copyright: the author(s), 2020 licensee pagepress, italy advances in oceanography and limnology, 2020; 11:8743 doi: 10.4081/aiol.2020.8743 references abdel-mawgoud am, aboulwafa mm, hassouna nah, 2008. characterization of surfactin produced by bacillus subtilis isolate bs5. appl. biochem. biotechnol. 150:289-303. arjmand mm, rezaee a, nasseri s, eshraghi ss, 2016. removing bacillus subtilis spores from drinking water using a bipolar electrochemical method. int. j. electrochem. sc. 11:10080-10086. boone dr, liu y, zhao zj, balkwill dl, drake gr, stevens to, aldrich hc, 1995. bacillus infernus sp. nov., an fe(iii)and mn(iv)-reducing anaerobe from the deep terrestrial subsurface. int. j. systematic bacteriol. 45:441-448 dias ac, andreote fd, dini-andreote f, lacava pt, sá alb, melo is, azevedo jl, araújo wl 2009. diversity and biotechnological potential of culturable bacteria from brazilian mangrove sediment. world. j. microbiol. biotechnol. 25:1305-1311. donio mbs, ronica sfa, viji vt, velmurugan s, jenifer ja, michaelbabu m, citarasu t, 2013. isolation and characterization of halophilic bacillus sp. bs3 able to produce pharmacologically important biosurfactants. asian. pac. j. trop. med. 6:876-883. dunlap ca, kwon sw, rooney ap, kim sj, 2015. bacillus paralicheniformis sp. nov., isolated from fermented soybean paste. int. j. syst. evol. micr. 65:3487-3492. dunlap ca, schisler da, perry eb, connor n, cohan fm, rooney ap, 2017. bacillus swezeyi sp. nov. and bacillus haynesii sp. nov., isolated from desert soil. int. j. sys. evol. micr. 67:2720-2725. fan b, blom j, klenk hp, borriss r, 2017. bacillus amyloliquefaciens, bacillus velezensis, and bacillus no nco mm er cia l u se on ly phylogeny of bacterial community of mangrove forest sediments 9 siamensis form an “operational group b. amyloliquefaciens” within the b. subtilis species complex. front. microbiol. 8:22. farrar we jr, 1963. serious infection due to “non-pathogenic” organisms of the genus bacillus. review of their status as pathogens. am. j. med. 34:134-141. feinberg l, jorgensen j, haselton a, pitt a, rudner r, margulis l, 1999. arthromitus (bacillus cereus) symbionts in the cockroach blaberus giganteus: dietary influences on bacterial development and population density. symbiosis 27:109-123. fry jc, 2004. culture-dependent microbiology, p. 80-87. in: a.t. bull (ed.), microbial diversity and bioprospecting. american society of microbiology, washington, dc. ghosh a, dey n, bera a, tiwari a, sathyaniranjan kb, chakrabarti k, chattopadhyay d, 2010. culture independent molecular analysis of bacterial communities in the mangrove sediment of sundarban. india. saline systems 6:1-11. guixa-boixareu n, calderón-paz j, heldal m, bratbak g, pedrós-alió c, 1996. viral lysis and bacterivory as prokaryotic loss factors along a salinity gradient. aquat. microb. ecol. 11:215-227. holguin g, gonzalez-zamorano p, de-bashan le, mendoza r, amador e, bashan y, 2006. mangrove health in an arid environment encroached by urban development a case study. sci. total. environ. 363:260-274. javor bj, 2012. hypersaline environments: microbiology and biogeochemistry. springer -verlag: 328 pp. jasim b, sreelakshmi s, mathew j, radhakrishnan ek, 2016. identification of endophytic bacillus mojavensis with highly specialized broad spectrum antibacterial activity. 3 biotech. 6:187. la jeon y, yang jj, kim mj, lim g, cho sy, park ts, suh jt, park yh, lee ms, kim sc, lee hj, 2012. combined bacillus licheniformis and bacillus subtilis infection in a patient with oesophageal perforation. j. med. microbiol. 61:1766-1769. janek t, łukaszewicz m, rezanka t, krasowska a, 2010. isolation and characterization of two new lipopeptide biosurfactants produced by pseudomonas fluorescens bd5 isolated from water from the arctic archipelago of svalbard. bioresour. technol. 101:6118-6123. jørgensen jn, laguna js, millán c, casabuena o, gracia mi, 2016. effects of a bacillus-based probiotic and dietary energy content on the performance and nutrient digestibility of wean to finish pigs. anim. feed sci. technol. 221:54-61. kaewtapee c, burbach k, tomforde g, hartinger t, camarinhasilva a, heinritz s, seifert j, wiltafsky m, mosenthin r, rosenfelder-kuon p, 2017. effect of bacillus subtilis and bacillus licheniformis supplementation in diets with lowand high-protein content on ileal crude protein and amino acid digestibility and intestinal microbiota composition of growing pigs. j. anim. sci. biotechnol. 8:37. kaur i, das ap, acharya m, klenk hp, sree a, mayilraj s, 2012. planococcus plakortidis sp. nov., isolated from the marine sponge plakortis simplex (schulze). int. j. syst. evol. micr. 62: 883-889. kiran gs, thomas ta, selvin j, sabarathnam b, lipton ap, 2010. optimization and characterization of a new lipopeptide biosurfactant produced by marine brevibacterium aureum msa13 in solid state culture. bioresour. technol. 101: 2389-2396. kumar s, stecher g, tamura k, 2016. mega7: molecular evolutionary genetics analysis version 7.0 for bigger datasets. mol. biol. evol. 33: 1870-1874. li l, mendis n, trigui h, oliver jd, faucher sp, 2014. the importance of the viable but non-culturable state in human bacterial pathogens. front. microbiol. 5:258. lane dj, 1991. 16s/23s rrna sequencing, p. 115-175. in: e. stackebrandt and m. goodfellow (eds.), nucleic acid techniques in bacterial systematics. wiley & sons, chichester. larkin ma, blackshields g, brown np, chenna r, mcgettigan pa, mcwilliam h, valentin f, wallace im, wilm a, lopez r, thompson jd, 2007. clustal w and clustal x version 2.0. bioinformatics 23:2947-2948. larsen n, thorsen l, kpikpi en, stuer-lauridsen b, cantor md, nielsen b, brockmann e, derkx pm, jespersen l, 2014. characterization of bacillus spp. strains for use as probiotic additives in pig feed. appl. microbiol. biotechnol. 98:1105-1118. lee sy, dunn rjk, young ra, connolly rm, dale per, dehayr r, lemckert cj, mckinnon s, powell b, teasdale pr, welsh dt, 2006. impact of urbanization on coastal wetland structure and function. austral. ecol. 31:149-163. liang jb, chen yq, lan cy, tam nf, zan qj, huang ln, 2007. recovery of novel bacterial diversity from mangrove sediment. mar. biol. 150:739-747. ma c, zhou s, lu q, yang g, wang d, zhuang l, li f, lei f, 2013. decolorization of orange i under alkaline and anaerobic conditions by a newly isolated humus-reducing bacterium, planococcus sp. mc01. int. biodeter. biodegr. 8:17-24. lima tm, fonseca af, leão ba, mounteer ah, tótola mr, borges ac, 2011. oil recovery from fuel oil storage tank sludge using biosurfactants. j. bioremed. biodegrad. 2:104172. mandell gl, bennett je, dolin r, 2010. mandel, douglas and bennett’s principles and practice of infectious diseases, 7th ed. churchill livingstone, philadelphia. martínez cruz p, ibáñez al, monroy hermosillo oa, ramírez saad hc, 2012. use of probiotics in aquaculture. isrn microbiology 2012;2012:916845. meidong r, doolgindachbaporn s, jamjan w, sakai k, tashiro y, okugawa y, tongpim s, 2017. a novel probiotic bacillus siamensis b44v isolated from thai pickled vegetables (phak-dong) for potential use as a feed supplement in aquaculture. j. gen. appl. microbiol. 63:246-253. meng qw, yan l, ao x, zhou tx, wang jp, lee jh, kim ih, 2010. influence of probiotics in different energy and nutrient density diets on growth performance, nutrient digestibility, meat quality, and blood characteristics in growing-finishing pigs. j. anim. sci. 8:3320-3326. mondal kc, pati br, 2000. studies on the extracellular tannase from newly isolated bacillus licheniformis kbr 6. j. basic microbiol. 4:223-232. nakagawa y, sakane t, yokota a, 1996. emendation of the genus planococcus and transfer of flavobacterium okeanokoites zobell and upham 1944 to the genus planococcus as planococcus okeanokoites comb. nov. int. j. syst. evol. microbiol. 46: 866-870. nicholson wl, 2002. roles of bacillus endospores in the environment. cell. mol. life. sci. 59:410-416. no nco mm er cia l u se on ly p. javid et al.10 pacwa-plociniczak m, plaza ga, piotrowska-seget z, cameotra ss, 2011. environmental applications of biosurfactants: recent advances. int. j. mol. sci. 12:633-654. palmisano mm, nakamura lk, duncan ke, istock ca, cohan fm, 2001. bacillus sonorensis sp. nov., a close relative of bacillus licheniformis, isolated from soil in the sonoran desert, arizona. int. j. syst. evol. micr. 51:1671-1679. por fd, 1980. a classification of hypersaline waters, based on trophic criteria. mar. ecol. 1:121-131. ruiz-garcia c, bejar v, martinez-checa f, llamas i, quesada e, 2005. bacillus velezensis sp. nov., a surfactant-producing bacterium isolated from the river velez in malaga, southern spain. int. j. syst. evol. microbiol. 55: 191-195. salehipour milani a, jafar beglu m, 2012. satellite-based assessment of the area and changes in the mangrove ecosystem of the qeshm island, iran. j. environ. res. dev. 7:1052–1060. stenfors arnesen lp, fagerlund a, granum pe, 2008. from soil to gut: bacillus cereus and its food poisoning toxins. fems microbiol. rev. 32:579-606. stoica c, sorescu i, 2017. abis online advanced bacterial identification software, an original tool for phenotypic bacterial identification. regnum prokaryotae 10:1-13. available from http://www.tgw1916.net/bacteria_logare_desktop.html thu pm, populus j, 2007. status and changes of mangrove forest in mekong delta: case study in tra vinh, vietnam. estuar. coast. shelf sci. 71:98-109. thwaite je, atkins hs, 2012. bacillus: anthrax; food poisoning, p. 237-244. in: d. greenwood, m. barer, r. slack and w. irving (eds.), medical microbiology. elsevier, amsterdam. valentine dl, 2007. adaptations to energy stress dictate the ecology and evolution of the archaea. nat. rev. microbiol. 5:316-323. vasavada sh, thumar jt, singh sp, 2006. secretion of a potent antibiotic by salt-tolerant and alkaliphilic actinomycete streptomyces sannanensis strain rjt-1. curr. sci. 91:13931397. ventosa a, 2006. unusual micro-organisms from unusual habitats: hypersaline environments, p. 223-254. in: n.a. logan, h.m. lappin-scott and p.c.f. oyston (eds.), proceedings sgm symposium 66: prokaryotic diversity – mechanisms and significance. cambridge university press, cambridge;. wei x, ji z, chen s. 2010. isolation of halotolerant bacillus licheniformis wx-02 and regulatory effects of sodium chloride on yield and molecular sizes of poly-γ-glutamic acid. appl. biochem. biotechnol. 160:1332-1340. yan l, guizhu c, 2007. physiological adaptability of three mangrove species to salt stress. acta ecol. sinica 27:22082214. yeon sh, jeong wj, park js, 2005. the diversity of culturable organotrophic bacteria from local solar salterns. j. microbiol. 43:1-10. yoon jh, weiss n, kang kh, oh tk, park yh, 2003. planococcus maritimus sp. nov., isolated from sea water of a tidal flat in korea. int. j. syst. evol. microbiol. 53:20132017. yoon yj, im kh, koh yh, kim sk, kim jw, 2003. genotyping of six pathogenic vibrio species based on rflp of 16s rdnas for rapid identification. j. microbiol. 41 312-319. zahed ma, rouhani f, mohajeri s, bateni f, mohajeri l, 2010. an overview of iranian mangrove ecosystems, northern part of the persian gulf and oman sea. acta ecol. sinica 30:240-244. zeigler dr, perkins jb, 2015. the genus bacillus, p. 309-337. in: e. goldman and l.h. green 2nd (eds.), practical handbook of microbiology. crc press, boca raton. zhang x, gao j, zhao f, zhao y, zhanshuang l, 2014. characterization of a salt-tolerant bacterium bacillus sp. from a membrane bioreactor for saline wastewater treatment. j. environ. sci. 26:1369-1374. no nco mm er cia l u se on ly layout 1 introduction the search of factors shaping the protistan plankton community has been extended and enhanced in the last years thanks to the combination of molecular data with tools from macroecology. studies on habitat heterogeneity in different marine ecosystems (grossmann et al., 2016; massana et al., 2015; de vargas et al., 2015; wu et al., 2020) have shown that plankton microeukaryotes present different community compositions depending both on environmental factors, affecting the habitat features, and on geographical distance on the small scale (horner-devine et al., 2004; martiny et al., 2006; wu et al., 2020). habitat heterogeneity is particularly marked in transitional water ecosystems (twe; european union, 2000; elliot and mclusky, 2002), such as coastal lagoons, due to their geomorphology, the close benthic–pelagic coupling, the freshwater inputs, and the connection with the adjacent marine ecosystems through tides and currents, (basset et al., 2006; mclusky and elliot, 2007; vadrucci et al., 2007). these twe features have a profound effect on plankton composition and distribution at both spatial and temporal scales (cloern and jassby, 2010). the lagoon of venice (lov) is the largest italian lagoon. this twe belongs to the long term ecosystem research (lter) national (lter-italy), european (ltereurope) and global (ilter) networks, which are essential components of the worldwide efforts to improve our knowledge of the structure and functions of ecosystems and of their long-term response to environmental, societal, and economic drivers (mirtl et al., 2018). they are distributed networks of research sites for multiple purposes in the fields of ecosystem, critical zone, and socio-ecological research. lter-italy consists of 79 research sites, which include terrestrial, freshwater, transitional and marine ecosystems (pugnetti et al., 2013; bergami et al., 2018). for what concerns planktonic protists of the lov, past lter activities addressed almost exclusively phytoplankton communities, for which a consistent knowledge has been gathered through morphology-based studies (bernardi aubry et al., 2013, 2017). the lov community includes taxa that are representative of the complex and open system of the lagoon itself, from the coastal sea to the rivers and the benthic-pelagic system: overall, they show a rather even distribution across distinct lagoon sites, while habitat heterogeneity could barely be detected (bernardi aubry et al., 2013). high throughput sequencing (hts) metabarcoding has so far been applied to the lov bacterial communities (quero et al., 2017) and, more recently, also to the whole protistan community, in a comparison between article spatial diversity of planktonic protists in the lagoon of venice (lter-italy) based on 18s rdna simona armeli minicante,1 roberta piredda,2 stefania finotto,1 fabrizio bernardi aubry,1* francesco acri,1 alessandra pugnetti,1 adriana zingone2 1institute of marine sciences, national research council (ismar-cnr), venice; 2department of integrative marine ecology, stazione zoologica anton dohrn, naples, italy abstract transitional waters are subject to a high degree of variability in space and time. in this study, protist plankton communities of the lagoon of venice were compared among four sites characterised by different environmental conditions with a metabarcoding approach. high throughput sequencing (hts) of the v4-18s rdna fragment in 32 samples collected on four dates, from april 2016 to february 2017, produced 1,137,113 reads, which were grouped into 4058 otus at 97% similarity. bacillariophyta and ciliophora were the most abundant groups in the entire dataset in terms of read number (27.6% and 16.6%, respectively), followed by dinophyta (10.9%), cryptophyceae (9.7%), and syndiniales (6.1%). the contribution of protist groups markedly varied across the seasons, but spatial differences were also recorded in the lagoon. in april, a higher contribution of bacillariophyta characterized st1 and 5 (68.0% and 61.1%), whereas sts2 and 3 showed a higher percentage of ciliophora (18.6 and 23.4%, respectively) and dinoflagellates (10.3 and 7.7%). in july, diatom blooms occurred at sts1, 2 and 3, with some differences in the dominant species. at st2 dinophyta reached the highest contribution of the whole sampling period in the area (30.6%), while st5 was quite distinct, with a low contribution of diatoms and a dominance of ciliophora (34.0%) and trebouxiophyceae (36.4%). all the stations in november were characterized by relatively high abundance of ciliophora (21.4-51.9%). in february, diatom contribution was relevant only at st5 (29.3%), teleaulax acuta peaked at st3 (ca. 36%), syndiniales at st2 (38.8%) and dictyochophyceae at st1 (24.2%). the α-diversity indexes (observed otus, shannon and pielou evenness) showed a high variability over space and time. diversity and community composition were rather similar between the intermediate and deeper sts2 and 3 on all sampling dates whereas they at time differed between the landward and shallow sts1 and 5. while the most marked differences occurred over the temporal scale, the depth of the station and the relatedness with the external marine coastal environment appear to play a major role in the spatial distribution of protist communities within the lagoon. overall, the quite stable spatial differences in spite of the significant seasonal changes reflected the hydro-geological heterogeneity of the sampling stations, indicating a major influence of the landsea gradient in the lagoon. no nco mm er cia l u se on ly s. armeli minicante et al.36 the lagoon and the adjacent marine waters of the gulf of venice (armeli minicante et al., 2019). in spite of potential pitfalls of the metabarcoding approach, which can poorly resolve the diversity of some taxa (piredda et al. 2018) or miss them altogether (massana et al., 2015), the latter study based on the v4-18s rrna metabarcodes largely increased the diversity knowledge not only for protists that have traditionally been neglected (i.e., heterotrophs, parasites, picoeukaryotes and other featureless groups), but also for the main phytoplankton taxa studied in the long term with morphology-based approaches (i.e., diatoms and dinoflagellates). in addition, hts results highlighted profound differences in the structure of the protistan communities between the lagoon and the external coastal waters of the gulf of venice: heterogeneity appears strong enough to allow for ecological segregation in the two environments, despite no clear barrier to dispersal processes among local protist communities. in this study, we explore the effects of habitat heterogeneity within the lov by comparing protistan communities among four individual lov sites that were lumped together in the previous hts study. the main goal of this work is to test whether the overall picture of rather spatially homogeneous lov communities emerging from morphologically-based studies on phytoplankton is confirmed by the higher resolution and better coverage allowed by the hts-metabarcoding approach. we aim at assessing if environmental selection plays a more important role than dispersal in shaping the microeukaryotic communities within the lov and identifying the most relevant local conditions that could affect them. methods study area the lov (northern adriatic sea, mediterranean sea) is a microtidal, polyhaline lagoon classified as a transitional water body (european union, 2000) and is the largest wetland (550 km2) in the mediterranean sea (poggioli, 2008). this delicate and fragile ecosystem is surrounded by densely inhabited and industrial areas and is affected by high numbers of tourists, as well as intensive fisheries and aquaculture activities. it has an average depth of 1 m and is morphologically characterized by the presence of large shallow areas and a network of deeper (5–10 m) channels. three inlets interrupt the sandbars separating the lagoon from the sea, allowing water turnover during tidal cycles, whose amplitude is 100 cm with maxima of 150 cm. water residence time, resulting from the interactions of tide, wind and topography, ranges from a few days close to the inlets to one month in landward areas (umgiesser et al., 2014; ghezzo et al., 2015). twelve main tributaries discharge an annual average of about 35 m3 s−1 of freshwater into the lagoon, with seasonal peaks in spring and autumn. in recent years aquatic angiosperms have recolonized the lagoon, especially in the central and northern basins (sfriso and buosi, 2018); however, in large areas the main primary producer is phytoplankton (acri et al., 2004, sfriso et al., 2005; bernardi aubry et al., 2013). in this study four sampling campaigns were conducted in april, july, november 2016 and february 2017, respectively, at four stations (st1, st2, st3 and st5, fig.1) that are regularly sampled for phytoplankton and abiotic factors within lter activities (bernardi aubry et al., 2013). the stations are all influenced, at different extents, by marine and freshwater inputs: st1 (s. giuliano, 2 m depth ) is in an area of intense maritime traffic (bianchi et al. 1996, socal et al. 1999) which collects urban waste from the town of mestre; st2 (marghera, 9.5 m) is affected by industrial pollution (perin 1975, guerzoni et al. 2007); st3 (fusina, 3.5 m) is influenced by heat emissions from the porto marghera power station (alberighi et al., 1992, socal et al., 1999); st5 (palude della rosa, depth 2.7 m) is a typical inland marshy lagoon area (bianchi et al., 1999). differences among the stations mainly depend on the varying degrees of impact of the interactions among the adjacent marine waters, riverine inputs and benthic-pelagic coupling (bernardi aubry et al., 2013). abiotic parameters for each station, temperature and salinity were measured with a ctd sbe 911 probe and water samples were collected at the near-surface layer using a niskin bottle. for chlorophyll a (chl a), 500 ml of water samples were immediately filtered through whatman gf/f fibreglass filters (nominal porosity = 0.7 µm), which were stored frozen and subsequently analyzed according to holm-hansen et al. (1965). dissolved inorganic nitrogen (din), as sum of ammonium (n-nh4), nitrites (n-no2) and nitrates (n-no3), was analyzed along with orthophosphates (p-po4) and orthosilicates (si-sio4) with a systea easychem plus according to grasshoff et al. (1983). filtration, dna extraction and sequencing at each station, 3 l of seawater were prefiltered on a 200 μm mesh-size net and then filtered onto cellulose ester 1.2 μm pore size filters (47 mm ø, whatman) using a peristaltic pump. for each station a duplicate was performed, obtaining 32 filters that were stored at -80°c until molecular analysis. total dna from each filter was extracted using the dneasy 96 plant kit (qiagen) and concentrations were determined with the qubit dsdna hs kit (thermofisher). dna samples were stored at −80°c until pcr. the hypervariable v4 region of eukaryno nco mm er cia l u se on ly planktonic protist diversity in the lagoon of venice 37 ote ssu rdna gene was amplified and sequenced (2 x 250 bp sequencing) on an illumina miseq platform as described in piredda et al. (2017). sequence analyses paired-end reads were processed using mothur v.1.33.0 (schloss et al., 2009). contigs between read pairs were assembled and differences in base calls in the overlapping region were solved using δq parameter as described in kozich et al. (2013). primer sequences were removed (pdiffs=3) and no ambiguous bases were allowed; the maximum homopolymer size was 8 bp. the remaining sequences were de-replicated and screened for chimeras using uchime in de novo mode (edgar et al., 2011). sequences were clustered into operational taxonomic units (otus) at 97% of similarity using vsearch (rognes et al., 2016) clustering (method=dgc) through mothur. otus containing only one read (singleton) were removed from downstream analyses. taxonomic assignment was performed on a single representative sequence from each otu (the most abundant) using blastn (altschul et al., 1990) against the pr2 database (v.4.10.0; https://doi.org/10.6084/m9. figshare.5913181; guillou et al., 2013), discarding the assignations with similarity ≤90% and query coverage ≤70% of the sequence length. statistical analyses the vegan r package (r core team, 2014; oksanen et al., 2016) was used for multivariate analyses. the dataset was normalized with a random subsampling to the second-lowest number of sequences (n = 74,396) with the ‘rrarefy’ function. observed otus and α-diversity estimators (shannon diversity and pielou evenness) were calculated and non-metric multidimensional scaling (nmds) was performed using the ‘metamds’ function (vegan) based on a bray-curtis dissimilarity matrix. permanova (adonis function in vegan) was used to detect significant association between environmental parameters and protist community structure. the selected variables were used to perform canonical correspondence analysis (cca). results abiotic parameters and chlorophyll a abiotic parameters and chl a showed a wide variability over time and among stations (tab. 1). salinity values (21-32.3) most of the times were lowest at sts1 and 5 (>21 and 23, respectively), while they were above 30 all year round at the other two stations. temperature ranged between 9.4°c (st2, february) and 31.5°c (st3, july), with fig. 1. study area with the four sampling stations. no nco mm er cia l u se on ly s. armeli minicante et al.38 marked differences (6-10°c) among stations on the same sampling date, and st3 always showing the highest values. din and silicate concentrations peaked at all the stations, and particularly at sts1 and 5, in november (25.4-88.1 µm) and february (9.1-61.7 µm), the maxima occurred mostly at st1 in every date. phosphates were always highest at st1 (0.6-3 µm), while at the other stations they fluctuated between 0,1 and 1.1 µm without a clear seasonal or spatial pattern. chl a concentrations were higher at sts1 and 5 (2.7 and 2.2 µgl–1, respectively) in april, then they peaked in july at sts1, 2 and 5 (12.6, 11.8 and 9.6 µgl–1, respectively) and were relatively low at all stations in november (0.20.7 µgl–1) and february (0.1-0.7 µgl–1). the highest chl a concentrations were recorded at st1 on all dates but february, when the maximum was at st5. protist diversity and seasonality the total cleaned dataset contained 3,588,416 reads from 32 samples. duplicate samples were generally coupled in hierarchical clustering based on bray-curtis distance (data not shown) and were hence pooled into 16 samples. after the normalization, the dataset consisted of 1,137,113 v4-18s reads (supplementary material 1), clustering into 4,058 otus mostly belonging to 34 highlevel taxonomic groups (fig. 2). bacillariophyta and ciliophora were the most abundant groups in terms of read number (27.6% and 16.6%, respectively), followed by dinophyta (10.9%), cryptophyceae (9.7%), syndiniales (6.1%). chrysophyceae, dictyochophyceae, mamiellophyceae, trebouxiophyceae, mast (nanoheterotrophic marine stramenopiles) and chlorophyceae ranged between 2 and 5%. the remaining 23 taxonomic groups were included in “other eukaryotes” (8.9%). over the four sampling dates, a change in the proportion of the main groups was evident at all stations, which at times also differed among them (fig. 3). common to sts1, 2 and 3 was the high relative abundance of bacillariophyta both in april (max 67.9%, at st1) and july (max 87%, at st3) and their very low abundance in november (1.1%, at st2) and february (1.9%, at st2). at st5, baciltab. 1. environmental variables at the four stations of the lagoon of venice investigated on four sampling dates. month station temperature salinity din si-sio4 p-po4 chl-a (°c) (µm) (µm) (µm) (µgl–1) april st1 14.1 20.9 30.3 36.7 0.5 2.7 st2 17.3 30.1 17.9 13.1 0.5 0.8 st3 24.3 29.5 23.1 18.3 0.3 0.1 st5 14.1 30.3 13.9 10.8 0.1 2.2 july st1 25.3 29.5 15.2 41.6 2.3 12.5 st2 28.8 31.3 6.5 13.6 0.2 11.7 st3 31.5 30.0 17.5 16.6 1.2 1.7 st5 25.5 27.1 12.3 43.8 0.6 9.5 november st1 14.5 28.3 83.9 88.1 2.5 0.7 st2 14.7 31.0 36.9 28.4 1.0 0.3 st3 22.1 31.6 36.0 25.3 0.8 0.2 st5 14.0 25.9 63.8 37.2 0.4 0.5 february st1 10.4 21.0 75.8 61.6 2.9 0.5 st2 9.4 31.3 53.8 27.1 0.7 0.1 st3 15.6 32.9 27.4 12.7 0.4 0.5 st5 11.8 23.3 24.4 9.1 0.2 0.7 fig. 2. taxonomic diversity of the normalized dataset as percentage of reads belonging to the main 10 high-level taxonomic groups and to the broad group “other eukaryotes” which includes: apusozoa, radiozoa, amoebozoa, stramenopiles, apicomplexa, perkinsea, palagophyceae, bolidophyceae, telonemia, pirsonia, bicoecea, moch, prymnesiophyceae, centroheliozoa, katablepharidaceae, choanoflagellatea, labyrinthulea, prasinophyceae, oomycota, chlorodendrophyceae, chlorophyceae, cercozoa, rhodophyta, picozoa. no nco mm er cia l u se on ly planktonic protist diversity in the lagoon of venice 39 lariophyta contribution was highest in april (61.1%) too, but low in july (10.9%) and relatively high again in february (29.3%). other seasonal patterns shared among stations concerned the higher contribution of trebouxiophyceae in july (up to 23.6% and 5 36.4% at sts1 and 5, respectively), of ciliophora at all stations in november (21.3-51.9%), and the low contribution of cryptophyceae and mamiellophyceae in july. besides, at all stations dictyochophyceae were more abundant at almost all stations in february (<24.1% at st1), while the contribution of chrysophyceae was significant in november, especially at st5 (12.9%). dinophyta attained more than the 20% of the total abundance only at st2 in july (30.6%) and november (23.7%). within each sampling date, some dissimilarities among stations were also evident. in april, the main difference was the highest contribution of bacillariophyta at sts1 and 5, which also showed the highest chl a concentrations. yet, diatom species composition differed between these two stations: thalassiosira concaviuscula dominated both the group and the whole protist assemblage at st1 (56.6% of total reads), while it covered less than 1% at st5, where the dominant species (50%) was cyclotella sp. at the other two stations, diatoms were less abundant and mainly represented by thalassiosira spp., chaetoceros tenuissimus, navicula and skeletonema. sts2 and 3 differed from sts1 and 5 also for the highest abundance of ciliophora (18.6% and 23.4%), mainly represented by strombidida and by the tintinnid tintinnopsis, and of dinoflagellates (<10%), with the autotrophic species gymnodinium dorsalisulcum and heterocapsa pygmaea. mamiellophyceae, mainly ostreococcus mediterraneus, were most abundant at st3 (10.9%), cryptophyceae at st2 (15.0%) with teleaulax acuta and rhodomonas sp. in july a diatom bloom occurred at sts1, 2 and 3, with the highest relative percentage contribution (87.5 %) at st. 3 and the lowest (50.7%) at st1. chaetoceros tenuissimus dominated at sts1 and 3, the cymatosiracean minutocellus polymorphus at st2. there, dinophyta, with g. dorsalisulcum and h. pygmaea (92.0%), attained the highest contribution (30.6%) of the whole sampling period. st5 was quite distinct from the others due to a much lower percentage of diatoms (10.9%) and the dominance of ciliophora (34%, mainly parastrombidinopsis sp.) and trebouxiophyceae (36.4%), the latter mainly represented by the green coccoid picochlorum which was also abundant at st1 (23.6%). in november all the stations were characterized by the high abundance and diversity of ciliophora (21.351.9%), among which the most abundant were an unknown strobilidiidae, rimostrombidium and strombidium. in addition to ciliophora, sts2 and 3 showed a quite similar composition in terms of both groups and species, with a high proportion of dinophyta (23.7% at st2, 15.1% at st3), mainly represented by h. pygmaea, the parasitic group syndiniales (13.3 at st2, 16.1% at st3), heterotrophs, such as mast (6.7% at st2, 4.4% at st3) and other eukaryotes (such as picozoa and other non-photosynthetic stramenopiles, 8.9% and 9.3% fig. 3. composition of the protist community at the four stations at each sampling date, based on the normalized dataset. no nco mm er cia l u se on ly s. armeli minicante et al.40 at sts 2 and 3, respectively ). cryptophyceae (with a high proportion of teleaulax acuta) peaked at st5 (20.5%) together with the dictyochophyceae pedinellales (18.7%). in february st5 differed from all the others for the dominance of diatoms (29.3%) mainly represented by benthic taxa, either marine (undatella and achnantes) or freshwater (cymbella), a higher percentage of ciliates (15.0%) and the unicellular rhodophyta (rhodella sp.) as the most abundant taxon (10.9%). the dominance of teleaulax acuta was remarkable at st3 (36%), while syndiniales were abundant at st2 (38.8 %) and dictyochophyceae at st1 (24.2%, mainly apedinella radians). dinophyceae at st2 and st3 were mainly represented by h. pygmaea. trends of alpha diversity (fig. 4) showed a great variability at both time and space scales. the effect of the diatom blooms was more evident in july, resulting in the lowest diversity values particularly at sts 2 and 3. in february, observed otu number was the lowest at st5, which however showed the highest evenness value. overall, diversity indexes were often similar within the two couples of stations, st1-st5 and st2-st3. the nmds ordination highlighted a relatively regular pattern in relation to the spatial and temporal scales (fig. 5). nmds2 values below or above zero split sts2 and 3 (bottom of the plot) from sts1 and 5 (top), whereas nmds1 separated february and november communities (to the left) from april and july ones (to the right), the latter being the most peculiar ones at all four sites. sts2 and 3 were strictly coupled in all sampling events but july, whereas sts1 and 5 were slightly more distant but still coupled in all sampling but february. permanova selected three environmental variables significantly correlated with the community composition: din, salinity, and temperature. in the cca (fig. 6) performed using these variables, the first two canonical axes only explained 23% of the total variance. yet, the analysis confirmed the clear separation of the july community, which showed a high correlation with temperature. february and november communities were associated with high din values. sts2 and 3 were coupled in all seasons, although at times more distant than in the nmds, whereas sts1 or 5 were often closer to one of the other two stations than between them. discussion high spatial and temporal variability is the hallmark for coastal environments and especially in twes, which are sensitive to an array of external drivers and pressures. in particular, in the lov the low average depth causes an accelerated exchange and biogeochemical cycling between the sediments and the water column (solidoro et al., 2010), which interacts with the variable fig. 4. trend of α-diversity indexes: observed otus, shannon diversity and pielou evenness. no nco mm er cia l u se on ly planktonic protist diversity in the lagoon of venice 41 nutrient load from land, tide cycle and meteo-climatic-driven hydrodynamism producing a complex space-time dynamic. to date the extent to which this complexity is reflected in the microbial community of the lagoon has only partially been understood also because of the limited resolution and partial coverage of the microbial diversity permitted by classical, microscopy-based approaches. in this study, the v4 -18s rrna metabarcode analysis allowed a deeper and more complete assessment of microeukariote communities of the area. although based on only four time points, the conditions recorded in this study can be considered quite indicative of the four seasons as they emerge from the previous knowledge of the area. the low plankton abundance during the autumn-winter time, the vegetative recovery in winter-spring and the summer peak match the “typical” seasonal trend observed for phytoplankton communities during 10 years of observations in the lov (bernardi aubry et al., 2013). the dominance of bacillariophyta in summer (and, of lesser importance, in spring) and their low contribution in autumn-winter, when our sampling may have missed the typical skeletonema peak, are also confirmed by light microscopy analysis (data not shown). while the dominance of tiny chaetoceros in spring–summer period was already known in the area (bernardi aubry et al., 2013), the importance of the cymatosyracean minutocellus polymorphus at st3 was not recognized before. the highest contribution of dinophyta in summer was already reported in literature but with lower abundance (average 2%; bernardi aubry et al., 2013). however, dinoflagellate abundance can be overestimated by hts studies because of the large amount of multicopy genes (piredda et al., 2017; gong et al., 2019). several species commonly found as abundant or featuring different seasons in this study were also reported as characterizing in others lagoons or estuarine environments, fig. 6. canonical correspondence analysis (cca) based on the normalized dataset, with the three environmental variables selected through permanova. fig. 5. non-metric mds ordination of protist community composition at the four stations, based on the bray-curtis dissimilarity matrix calculated on the otu table. stress value = 0.099. no nco mm er cia l u se on ly s. armeli minicante et al.42 such as teleaulax acuta, cyclotella sp. and minutocellus polymorphus (sarno et al., 1993; bérard-therriault et al., 1999; melo et al., 2010; guiry and guiry, 2019). from the diversity standpoint, spatial and temporal variations of heterotrophic, mixotrophic, parasitic such groups as ciliophora, mast and syndiniales, as well as of autotrophic and heterotrophic picoeukaryotes (e.g., trebouxiophyceae and picozoa) emerging in this study represent an absolute novelty and hint at quite intricate and dynamic trophic interactions within the microbial food web of the lagoon. for what concerns spatial differences, they appeared quite evident for the whole microeukaryotic communities compared to the overall homogeneity emerging from previous studies that only addressed phytoplankton. in those studies, despite some weak differences among areas, the most abundant species were shared by all stations investigated, whereby community composition appeared to be prevalently shaped by the general ecological characteristics of the lagoon ecosystem itself rather than by the specific characteristics of individual areas (acri et al., 2004; bernardi aubry et al., 2004, 2013). the latter picture is based on a much higher number of observations, but the higher resolution allowed by metabarcoding may depict more effectively differences and commonalities among sites. interestingly, some regularities in the spatial heterogeneity persisted across the sampling dates, i.e., in the coupling between sts1 and 5 and between sts2 and 3, where the structure of the protist communities probably reflects different hydro-geological characteristics. sts2 and 3 are characterized by greater depths (11 m and 4 m, respectively) compared to sts1 and 5 (1.5 m and 2.5 m) which, being the most landward sites, are more affected by resuspended (e.g., pennate diatoms) and/or freshwaters species from the mainland than by the incoming tide from the inlets. the rather regular coupling between sts2 and 3 may be explained by their position in the nearness of the oil tanker channel (canale dei petroli) that, connecting with the malamocco inlet, is more influenced by marine waters from the gulf of venice. this area is characterised by higher transit time and lower residence time respect to the other two sampling stations (guerzoni and tagliapietra, 2006). instead, sts1 and 5 are characterized by a low seabed and by a higher contribution of resuspended species (e.g., diatoms), and can be affected to varying degrees by the influence of the inland anthropogenic inputs (e.g., terrestrial waste in st1 or fresh water in st5), which are also very variable over time. this latter interpretation is confirmed by the higher abundance, at sts2 and 3, of taxa typical of the external marine environment such as cryptophyceae, dinophyta, syndiniales and mast (armeli minicante et al., 2019) with respect to sts1 and 5. the overall protistan community variance could be related to the average environmental conditions of the sampling stations, in turn depending on the site location in the lagoon, rather than to the instantaneous measured values of the environmental variables during each sampling event, which actually explained only a minor part of the variance (23%). in conclusion, this hts metabarcoding investigation reveals relatively stable spatial differences and substantial seasonal changes in the structure of the whole protistan communities of the lov, where habitat heterogeneity and connection coexist. protistan groups that have distinct functional roles (e.g., ciliophora, mast, syndiniales and picoeukaryotes), so far never investigated in the lov, also showed marked differences over the temporal and spatial scale, hinting at complex and dynamic relationships within the microbial community which warrant further investigations. overall some different hydro-geological characteristics that typify different areas contribute consistently in shaping the spatial microeukaryotic community configuration. in particular, the depth of the stations, which modulates the benthic-pelagic coupling, and the degree of relatedness with the external marine coastal environment or with the mainland, appear to play a major role in the spatial distribution of protist community within the lagoon. acknowledgments this work was possible thanks to partial financial support by italy-lifewatch-eric. rp was supported by the italian miur flagship project ritmare and the european union’s horizon 2020 research and innovation program embric (ga 654008). corresponding author: fabrizio.bernardi@ismar.cnr.it key words: 18s rdna; lagoon of venice; transitional environments; protists plankton diversity; habitat heterogeneity; lteritaly. received: 21 may 2020. accepted: 23 june 2020. this work is licensed under a creative commons attribution noncommercial 4.0 license (cc by-nc 4.0). ©copyright: the author(s), 2020 licensee pagepress, italy advances in oceanography and limnology, 2020; 11:8961 doi: 10.4081/aiol.2020.8961 references acri f, bernardi aubry f, bianchi f, boldrin a, comaschi a, rabitti s, socal g, 2004. changes in nutrients and plankton communities in the lagoon of venice from 1970s until today. j. mar. sys. 51: 321-329. no nco mm er cia l u se on ly planktonic protist diversity in the lagoon of venice 43 alberighi l, bianchi f, cioce f, socal g, 1992. [osservazioni durante un bloom di skeletonema costatum in prossimità della centrale termoelettrica enel di fusina porto-marghera (venezia)].[article in italian]. oebalia 17:321-322. armeli minicante s, piredda r, quero gm, finotto s, bernardi aubry f, bastianini m, pugnetti a, zingone a, 2019. habitat heterogeneity and connectivity: effects on the planktonic protist community structure at two adjacent coastal sites (the lagoon and the gulf of venice, northern adriatic sea, italy) revealed by metabarcoding. front. microbiol. 10:2736. doi: 10.3389/fmicb.2019.02736. basset a, sabetta l, fonnesu a, mouillot d, do chi t, viaroli p, giordani g, reizopoulou s, abbiati m, carrada gc, 2006. typology in mediterranean transitional waters: new challenges and perspectives. aquatic conserv: mar. freshw. ecosyst. 16:441-455. doi: 10.1002/aqc.767. bergami c, l’astorina a, pugnetti p, 2018. [i cammini della rete lter-italia il racconto dell’ecologia in cammino].[book in italian]. cnr edizioni, rome; 180 pp. bernardi aubry f, berton a, bastianini m, socal g, acri f, 2004. phytoplankton succession in a coastal area of the nw adriatic over a 10-years sampling period (1990-1999). continent. shelf res. 24:97-115. bernardi aubry f, acri f, bianchi f, pugnetti a, 2013. looking for patterns in the phytoplankton community of the mediterranean microtidal venice lagoon: evidence from ten years of observations. sci. mar. 7:47-70. doi: 10.3989/scimar. 03638.21. bernardi aubry f, pugnetti a, roselli l, stanca e, acri f, finotto s, basset a, 2017. phytoplankton morphological traits in a nutrient-enriched, turbulent mediterranean microtidal lagoon. j. plankton res. 39:564–576. doi: 10.1093/plankt/fbx008. cloern je, jassby ad, 2010. patterns and scales of phytoplankton variability in estuarine–coastal ecosystems. estuar. coast. 33: 230. https://doi.org/10.1007/s12237-009-9195-3. de vargas c, audic s, henry n, decelle j, mahé f, logares r., lara e, berney c, le bescot n, probert i, carmichaell m, poulain j, romac s, colin s, aury jm, bittner l, chaffron s, dunthorn m, engelen s, flegontova o, guidi l, horák a, jaillon o, lima-mendez g, lukeš j, malviya s, morard r, mulot m, scalco e, siano r, vincent f, zingone a, dimier c, picheral m, searson s, kandels-lewis s, tara oceans coordinators, acinas sg, bork p, bowler c, gorsky g, grimsley n, hingamp p, iudicone d, not f, ogata h, pesant s, raes j, sieracki me, speich s, stemmann l, sunagawa s, weissenbach j, wincker p, karsenti e, 2015. eukaryotic plankton diversity in the sunlit ocean. science 348:1261605. doi: 10.1126/science.1261605. edgar rc, haas bj, clemente jc, quince c, knight r. 2011, uchime improves sensitivity and speed of chimera detection. bioinformatics 27:2194-2200. doi: 10.1093/bioinformatics/btr381 elliott m, mclusky ds, 2002. the need for definitions in understanding estuaries. estuar. coast. shelf sci. 55: 815-827. doi: 10.1006/ecss.2002.1031. european union, 2000. common implementation strategy for the water framework directive (2000/60/ec) guidance document no 5 transitional and coastal waters – typology, reference conditions and classification systems. office for official publications of the european communities; luxembourg. ghezzo m, de pascalis f, umgiesser g, zemlys p, sigovini m, pérez-ruzafa cma, 2015. connectivity in three european coastal lagoons. estuaries coast. 38:1764-1781. doi: 10.1007/s12237-014-9908-0. gong w, marchetti a, 2019. estimation of 18s gene copy number in marine eukaryotic plankton using a next-generation sequencing approach. front. mar. sci. 6. doi: 10.3389/fmars. 2019.00219 grasshoff k, ehrhardt m., kremling k, 1983. methods of seawater analysis. wiley-vch verlag; pp. 419. doi: 10.1002/iroh.19850700232.. grossmann l, jensen m, heider d, jost s, glücksman e, hartikainen h, mahamdallie ss, gardner m, hoffmann d, bass d, boenigk j, 2016. protistan community analysis: key findings of a large-scale molecular sampling. isme j. 10:22692279. doi: 10.1038/ismej.2016.10. guerzoni s, tagliapietra d. 2006. [atlante della laguna. venezia tra terra e mare].[book in italian]. marsilio, venice: 242 pp. guerzoni s, rossigni p, sarretta a, raccanelli s, ferrari g, molinaroli e, 2007. pops in the lagoon of venice: budgets and pathways. chemosphere 67:1776-1785. holm-hansen o, lorenzen cj, holmes rw, strickland jdh, 1965. fluorometric determination of chlorophyll. j. conseil perm. int. explor. mer. 30:3-15. horner-devine m, lage m, hughes j, bohannan bjm, 2004. a taxa–area relationship for bacteria. nature 432:750–753. doi: 10.1038/nature03073. kozich jj, westcott sl, baxter nt, highlander sk, schloss pd, 2013. development of a dual-index sequencing strategy and curation pipeline for analyzing amplicon sequence data on the miseq illumina sequencing platform. appl. environ. microbiol. 79:5112-5120. doi: 10.1128/aem.01043-13 martiny j, bohannan b, brown j, colwell rk, fuhrman ja, green jl, horner-devine mc, kane m, krumins ja, kuske cr, morin pj, naeem s, ovreås l, reysenbach al, smith vh, staley jt, 2006. microbial biogeography: putting microorganisms on the map. nat. rev. microbiol. 4:102-112. doi: 10.1038/nrmicro1341. massana r, gobet a, audic s, bass d, bittner l, boutte c, chambouvet a, christen r, claverie jm, decelle j, dolan jr, dunthorn m, edvardsen b, forn i, forster d, guillou l, jaillon o, kooistra wh, logares r, mahé f, not f, ogata h, pawlowski j, pernice mc, probert i, romac s, richards t, santini s, shalchian-tabrizi k, siano r, simon n, stoeck t, vaulot d, zingone a, de vargas c, 2015. marine protist diversity in european coastal waters and sediments as revealed by high�throughput sequencing. environ. microbiol. 17:4035-4049. doi: 10.1111/1462-2920.12955 mclusky ds, elliott m, 2007. transitional waters: a new approach, semantics or just muddying the waters? estuar. coast. shelf sci. 71:359-363. doi: 10.1016/j.ecss.2006.08.025 mirtl m, borer e, djukic i, m.forsiusc, haubolda h, hugod w, jourdane j, lindenmayerf d, mcdowellg wh, muraokah h, orensteini de, pauwd jc, peterseila j, shibataj h, wohnera c, yuk x,haaseel p, 2018. genesis, goals and achievements of long�term ecological research at the global scale: a critical review of ilter and future directions. sci. tot. environ. 626:1439– 62. no nco mm er cia l u se on ly s. armeli minicante et al.44 oksanen j, blanchet fg, kindt r, legendre p, minchin pr, o’hara rb, simpson gl, solymos p, stevens mhh, szoecs e, wagner h, 2016. vegan: community ecology package, version 2.4-1. available from: https://cran. r-project. org/web/packages/vegan/index. html percopo i, ruggiero mv, balzano s, gourvil p, lundholm n, siano r, tammilehto a, vaulot d, sarno d, 2016. pseudo-nitzschia arctica sp. nov., a new cold-water cryptic pseudo-nitzschia species within the p. pseudodelicatissima complex. j. phycol. 52:184-99. doi: 10.1111/jpy.12395 perin g, 1975. [l’inquinamento chimico della laguna di venezia, p. 47-89]. in: cons. depur. acque della z. i. porto marghera (ed.), [problemi dell’inquinamento lagunare].[in italian]. venezia. piredda r, tomasino mp, d’erchia am, manzari c, pesole g, montresor m, kooistra wh, sarno d, zingone a, 2017. diversity and temporal patterns of planktonic protist assemblages at a mediterranean lter site. fems microbiol. ecol. 93. doi: 10.1093/femsec/fiw200 piredda r, claverie jm, decelle j, devargas c, dunthorn m, edvardsen b, eikrem w, forster d, kooistra whcf, logares r, massana r, montresor m, not f, ogata h, pawlowski j, romac s, sarno d, stoeck t, zingone a, 2018. diatom diversity through hts-metabarcoding in coastal european seas. sci. rep. 8:18059. doi: 10.1038/s41598-018-36345-9 power me, mills ls, 1995. the keystone cops meet in hilo. trends ecol. evol. 10:182-184. doi: 10.1016/s01695347(00)89047-3 pugnetti a, acri f, bernardi aubry f, camatti e, cecere e, facca c, franzoi p, keppel e, lugliè a, mistri m, munari c, padedda bm, petrocelli a, pranovi f, pulina s, satta ct, sechi n, sfriso a, sigovini m, tagliapietra d, torricelli p, 2013. the italian long-term ecosystem research (lter-italy) network: results, opportunities, and challenges for coastal transitional ecosystems. trans. wat. bull. 7:43-63. quero gm, perini l, pesole g, manzari c, lionetti c, bastianini m, marini m, luna gm, 2017. seasonal rather than spatial variability drives planktonic and benthic bacterial diversity in a microtidal lagoon and the adjacent open sea. mol. ecol. 26:5961-5973. doi: 10.1111/mec.14363. r core team, 2014. a language and environment for statistical computing. r foundation for statistical computing, vienna. solidoro c, bandelj v, bernardi aubry f, camatti e, ciavatta s, cossarini g, facca c, franzoi f, libralato s, melaku canu d, pastres r, pranovi f, raicevich s, socal g, sfriso a, sigovini m, tagliapietra d, torricelli p, 2010. response of the venice lagoon ecosystem to natural and anthropogenic pressures over the last 50 years, p. 483-511. in: m. kennish and h. paerl (eds.), coastal lagoons: critical habitats of environmental change. crc press, boca raton. schloss pd, westcott sl, ryabin t, hall jr, hartmann m, hollister eb, lesniewski ra, oakley bb, parks dh, robinson cj, sahl jw, stres b, thallinger gg, van horn dj, weber cf, 2009. introducing mothur: open-source, platform-independent, community-supported software for describing and comparing microbial communities. appl. environ. microbiol. 75:7537-7541. doi: 10.1128/aem.01541-09 sfriso a, facca c, ceoldo s, marcomini a, 2005. recording the occurrence of trophic level changes in the lagoon of venice over the ‘90s. environ. int. 31:993-1001.doi: 10.1016/j.envint.2005.05.009 sfriso a, buosi a, 2018. trophic status changes in the venice lagoon during the last 40 years. proceedings xlix sibm congress, genoa; p. 36-39. socal g, bianchi f, alberighi l, 1999.effects of thermal pollution and nutrient discharges on a spring phytoplankton bloom in the industrial area of the lagoon of venice. vie milieu 49:19-31. umgiesser g, ferrarin c, cucco a, de pascalis f, bellafiore d, ghezzo m, bajo m, 2014. comparative hydrodynamics of 10 mediterranean lagoons by means of numerical modelling. j. geophys. res. oceans. 119:2212-2226. doi: 10.1002/2013jc009512. vadrucci mr, cabrini m, basset a, 2007. biovolume determination of phytoplankton guilds in transitional water ecosystems of mediterranean ecoregion. transit. waters bull. 2: 83-102. doi: 10.1285/i1825229xv1n2p83 wu pf, li dx, kong lf, li yy, zhang h, xie zx, lin l, wang dz, 2020. the diversity and biogeography of microeukaryotes in the euphotic zone of the northwestern pacific ocean. sci. total environ. 698:134289 doi: 10.1016/j.scitotenv. 2019.134289 zingone a, d’alelio d, mazzocchi mg, montresor m, sarno d, lter-mc team, 2019. time series and beyond: multifaceted plankton research at a marine mediterranean lter site. nat. cons. 34:273-310. doi: 10.3897/nature conservation.34.30789. no nco mm er cia l u se on ly layout 1 advances in oceanography and limnology, 2015; 6(1/2): 2-12 original article doi: 10.4081/aiol.2015.5451 introduction biological invasions, i.e., the successful establishment of non-indigenous species (nis) in a given area, are long known to be one of the most serious threats to the conservation of the world biological diversity. in fact, nis are known to threat the survival of indigenous species, populations, and communities through hybridisation, competition, parasitism, predation, and the structural changes they cause to the colonised habitats (ehrenfeld, 2010; simberloff et al., 2013). furthermore, they are also known to cause substantial economic damages, and can be harmful for human health (pimentel et al., 2005, keller et al., 2011). although biological invasions are a pervasive global phenomenon which widely interests all the existing ecosystems, some evidences suggest that inland waters are especially prone to be invaded (gherardi, 2007; chandra and gerhardt, 2008). this is possibly due to the pronounced dispersal abilities of most of inland-water taxa (incagnone et al., 2015), to the naiveté of lakes and other inland water ecosystems to the effects of invaders owing to their evolutionary isolation (cox and lima, 2006), and, eventually, to the pivotal importance that these habitats have always had for the human civilisation. anthropogenic alterations of the pre-existing biocoenoses, both in terrestrial and marine environments, are in fact known to have likely facilitated the establishment of several opportunistic newcomers (chytrý et al., 2008; airoldi et al., 2015; see also boggero et al., 2014). moreover, freshwater ecosystems are globally experiencing the highest loss of biodiversity due to human activities (naiman and dudgeon, 2011). in spite of some early warnings (elton, 1958), and of the sound evidences of the impacts that non-indigenous species have on the indigenous biota, little efforts were paid to take a census and to monitor non-indigenous species in european inland waters till the end of the xx century; in some instances, confronting alien species was even suspected to be a form of xenophobia (simberloff, 2003). such a delay in approaching biological invasions a review on the animal xenodiversity in sicilian inland waters (italy) federico marrone,* luigi naselli-flores department of biological, chemical and pharmaceutical sciences and technologies, university of palermo, via archirafi 18, 90123, palermo, italy *corresponding author: federico.marrone@unipa.it abstract this paper reviews the available knowledge about faunal xenodiversity in sicilian inland waters (italy). the aim is to provide an updated checklist and bibliography of those non-indigenous species (nis) which occur in the island, and to identify possible threats to its native biological diversity. data were collected through an extensive literature search which encompassed also local journals, books, congress abstracts, and other grey literature. all the collected data were critically revised and, when possible, verified by consulting available collections or through dedicated sampling surveys. only those data contained in reports indicating precise occurrence localities, which were confirmed by our own observations and\or by at least two independent sources including at least a peer-reviewed publication, were considered as certain. data in literature that did not meet these criteria were considered doubtful and reported separately as unverified data. the information provided by websites has been excluded as it often contains unfounded and\or erroneous data. the fauna of sicilian inland waters host at present 31 confirmed nis. in addition, the presence of further 11 taxa is dubious. among the verified data, invertebrate and vertebrate taxa are nearly equally represented, with 15 and 16 taxa, respectively. with 16 species, the phylum chordata is by far the most represented, followed by mollusca (8 species) and arthropoda (6 species). most of these species were detected in the last 30 years due to the lack of previous regular studies on sicilian freshwaters. with few exceptions (e.g., the recent introduction of xenopus laevis, the african clawed frog), nis’ effects on native biota have not extensively studied in the island yet. although the top-down effects caused by introduced vertebrate taxa are known to deeply modify the native structure of the biota, little information is available on the impacts caused by invertebrate taxa, especially the microscopic ones. the presence in sicily of 11 nonnative species of bony fish is probably the most impacting threat to autochthonous fauna through predation, competition and hybridisation. the results shown in the paper highlight the importance and the urgency of more exhaustive investigations on nis in sicilian inland waters with special regard to less charismatic taxa whose effects on the native biota have never been evaluated yet. key words: biological invasions; mediterranean biodiversity; non-indigenous species; translocated species; parautochthonous taxa; allochthonous taxa. received: july 2015. accepted: september 2015. no nco mm er cia l u se on ly 3 f. marrone and l. naselli-flores interested even those countries with a longer tradition in limnological studies, and only recently national and international studies on invasion biology have been conducted. these have led to the creation of dedicated checklists and databases, available on the web, aimed at providing a complete census of the non-indigenous biota of the continent (e.g., daisie: delivering alien invasive species inventories for europe, www.europe-aliens.org, or aquatic alien species in german inland and coastal waters, http://www.aquatic-aliens.de/). however, the completion of an exhaustive census of the nis is intrinsically difficult given the dynamic nature of the phenomenon, and it is especially hard for those less-charismatic taxa and\or small-bodied organisms, which are more difficult to notice or to correctly identify. furthermore, the recent evidences that cryptic species or lineages often are the protagonists of widely overlooked cryptic invasions (saltonstall, 2002; marrone et al., 2011; van bocxlaer et al., 2015), stress the need for the implementation of molecular identification tools when dealing with biological invasions (blanchet, 2012). based on all these hindering factors, the known distribution of non-indigenous species often reflects the distribution of researchers interested in invasion biology and\or taxonomists, rather than that of the organisms themselves. such is the case of the non-indigenous biota of sicilian inland-waters, which was to date understudied, with sparse data, often published in scarcely accessible literature. as a consequence, even in the most recent reviews addressed to the non-indigenous biota of european and italian inland waters, sicily has often been considered jointly with sardinia (nocita and zerunian, 2007; tricarico et al., 2010; bianco, 2014); this approach reflects the paucity of data available for the two islands rather than any theoretical, biological, or historical reason. in some cases, sicilian nis were even not included in the analyses at all (marr et al., 2013; boggero et al., 2014). moreover, due to the aforementioned constraints, those few studies where sicily was considered as an independent region (gherardi et al., 2008) show rather fragmentary checklists. the lack of comprehensive data on nis makes their management and control difficult. actually, not all the introduced nis are able to successfully establish populations in the invaded areas, and even less prove to be actually invasive (i.e., noxious to the native biota see the tens rule, cf. williamson, 1996). however, when the invasiveness of a certain taxon becomes evident, it is often too late to control or eradicate it in spite of the efforts invested. accordingly, it has been suggested that the nis are to be considered guilty until proven otherwise, and that a quick and dirty response, aimed at eliminating the nis at their very first colonisation outset is strongly advisable and, possibly, the only way to solve the problem (gherardi, 2006). a detailed and timely monitoring of the current xenodiversity and of the ongoing biological invasions is thus needed when facing the challenge of protecting the indigenous species and ecosystems from biotic homogenization. the present paper collects and reviews all the available literature on the nis occurring in the fauna of sicilian inland waters, with the explicit aim to provide sound and updated baseline data for future studies and desirable management activities. methods definitions and study area in this paper we describe the animal xenodiversity, i.e., the diversity of the non-indigenous fauna, occurring in sicilian inland waters (cf. leppäkoski et al., 2002). inland waters are here defined according to the water framework directive (directive 2000/60/ec): all standing or flowing waters on the surface of the land. in contrast with some published papers dealing with italian xenodiversity (gherardi et al., 2008; tricarico et al., 2010; boggero et al., 2014), we are hereby considering nis (non-indigenous species) all those taxa which are occurring outside of their natural distribution range and dispersal potential, and which were introduced to sicily by human activities (iucn, 2000); we are therefore here including also the translocated species, i.e., those species which are native (autochthonous) to the italian peninsula or sardinia but that would be naturally absent in sicily. within the nis, we distinguish among those taxa introduced before the year 1500, called parautochthonous taxa, and those which were introduced in sicily after that date, i.e., the allochthonous taxa (gazzetta ufficiale della repubblica italiana, 2015). a further important partition is between those taxa which are present in sicily with self-sustaining breeding populations (established species) and those which are present on the island, but whose successful breeding was not observed and that might thus not be able to constitute self-sustaining populations; the non-occasional presence of the latter is due to an ongoing introduction of specimens in the wild (sporadic species). among the nis, we here consider invasive species those widespread non-indigenous species that have adverse effects on the invaded habitats (cf. gherardi, 2006). to date, no data on the occurrence of nis on the small-circum-sicilian islands are available; accordingly, the present paper focuses on the sicilian mainland only. bibliographical review a checklist of the nis reported to occur in sicily was compiled from an extensive literature search through journals, books, congress abstracts, and other grey literature. unfortunately, given the difficulties in exhaustively tracing no nco mm er cia l u se on ly 4alien species in sicilian inland water the grey literature, we cannot exclude that some information might have been missed. all collected data were critically revised and, when possible, checked by consulting available collections or through dedicated sampling surveys. in the frame of this paper, we considered as verified data those reports indicating precise occurrence localities, which were confirmed by our own observations and\or by at least two independent sources, including at least a peerreviewed publication. the information which did not meet these criteria was here considered doubtful and the taxa were reported separately as unverified. the information provided by websites has been excluded as it often contains anecdotic and\or erroneous data. results updated checklist and origin of the nis occurring in sicilian inland waters the updated checklist of the nis occurring in sicilian inland waters is reported in tabs. 1 to 3. overall, 31 nis were confirmed to be positively present in sicily (tabs. 1 and 2). another group includes 11 taxa whose presence in sicily is dubious and\or whose non-indigenous status in sicily is nowadays considered controversial (tab. 3). by taking into account only the verified data, invertebrates and vertebrates are nearly equally represented, with 15 and 16 taxa respectively. the phylum chordata is by far the most represented, with 16 species, followed by mollusca (8 species) and arthropoda (6 species) (fig. 1). overall, the commonest source for the nis in sicilian inland waters is the nearctic region, followed by the westand east-palaearctic subregions; only single taxa colonised sicily from the neotropical, afrotropical, oriental or australian regions (for more details see tabs. 1 and 2; fig. 2). in good accordance, the vast majority of the nis occurring in sicily are allochthonous taxa introduced after the xvi century, with only a single mammal species being a parautochthonous taxon (i.e., the brown rat), and three taxa with no data on their first introduction (i.e., the isopod proasellus banyulensis, and the fish perca fluviatilis and carassius auratus, cf. tabs. 1 and 2). unfortunately, nearly no information is available on how deliberate were the introductions of most of the nis in the inland waters of sicily, with a few exception as the one of gambusia holbrooki introduced in sicily between 1925 and 1927 to keep under control malaria-spreading mosquitoes (consoli, 1928; veronesi et al., 1997). however, a markedly different pattern is scored between invertebrate and vertebrate taxa. with the only exception of the brown rat, rattus norvegicus, all the remaining vertebrate taxa were likely the object of intentional introductions for ornamental, fishing, or sanitary purposes; conversely, the opposite pattern is scored among the invertebrates, which are mostly inconspicuous species unintentionally released in the wild along with intentionallyintroduced aquatic vertebrate species and\or ornamental plants (mazza et al., 2015). invertebrate nis and putative nis all the invertebrate nis listed in tab. 1 are known to be successfully established in sicily, being present with locally abundant, self-sustaining populations. among them, the oldest records are those referring to the gastropod haitia acuta, whose arrive in sicily dates back at least to the xix century (sowerby, 1873-1874), followed by the crustaceans daphnia ambigua, d. parvula, and proasellus banyulensis, all of them first collected in sicily in the ’80s (calvo et al., 1993; stoch et al., 1996; marrone et al., 2005). all the other taxa were reported to be present on the island only in the xxi century or are even still unpublished records (tab. 1), thus suggesting a recent significant increase in the rate of successful invasions by invertebrate nis in sicilian inland waters. among the putative nis whose presence in sicily is inferred from unverified data, the reports of the gastropod helisoma anceps and of the crustacean moina affinis are likely based on the misidentification of congener species known to occur on the island, while the report of the bivalve corbicula fluminea from western sicily is possibly to be ascribed to the mislabelling of a museum specimen (tab. 3). finally, the non-indigenous status for the gastropod galba truncatula in sicily is questioned by recent studies (liberto et al., 2010) which suggest that this taxon might in fact be autochthonous on the island. vertebrate nis and putative nis most of the vertebrate nis occurring in sicily are present with established, widespread populations (tab. 2). however, there are some exceptions: the rainbow trout, oncorhynchus mykiss, for instance, is a sporadic species, known to be unable to breed in sicilian inland waters. the presence of this species in the rivers of the island is to be ascribed to the ongoing release of specimens for recreational fishing. to date, no evidences of the presence of reproducing populations of the locally abundant red-eared slider, trachemys scripta elegans, in sicily are available, although the species is known to successfully breed in italy and might find in sicily suitable bioclimatic conditions for its reproduction (ficetola et al., 2009). finally, the case of emys orbicularis s.l. is quite peculiar as, although no pure populations of this non-indigenous species are known to occur on the island, there are evidences of the introduction of single e. orbicularis galloitalica fritz, 1995 specimens in sicilian localities inhabited by the endemic sicilian pond turtle, emys trinacris fritz, fattizzo, guicking, tripepi, pennisi, lenk, joger and wink, 2005 (lenk et al., 1999); no nco mm er cia l u se on ly 5 f. marrone and l. naselli-flores this has led to the genetic introgression of e. orbicularis genes in eastern-sicilian populations of e. trinacris (vamberger et al., 2015). the identity of some non-indigenous fish species reported to be present in sicily is still to be ascertained (tabs. 2 and 3); this is the case of the pike (esox cf. lucius), whose taxonomical identity has to be carefully checked in the light of the relatively recent description of the southern pike, esox cisalpinus bianco and delmastro, 2011, and of the chub (squalius cf. cephalus), for which tab. 1. list of the non-indigenous invertebrate animal species known to occur in sicilian inland waters. taxon category origin introduction status source(s) allochthonous vs intentional vs established vs parautochthonous unintentional sporadic nematoda secernentea spirurida anguillicolae anguillicola crassus kuwahara, niimi and itagaki, 1974 a southeast asia u e 4, 12 mollusca gastropoda hydrobiidae potamopyrgus antipodarum (j.e. gray, 1843) a new zealand u e 6, 12 lymnaeidae radix auricularia (linnaeus, 1758) a eurasia u e 6, 13 physidae haitia acuta (draparnaud, 1805) a north america u e 6, 9 planorbidae ferrissia fragilis (tryon, 1863) a north america u e 14 helisoma duryi (wetherby, 1879) a north america n.a. e 2, 11, 12 thiaridae melanoides tuberculata (o.f. müller, 1774) a tropical africa and asia n.a. e 9, 11 bivalvia dreissenidae dreissena polymorpha (pallas, 1771) a ponto-caspian region n.a. e 16 unionidae sinanodonta woodiana (lea, 1834) a east asia n.a. e 16 arthropoda crustacea branchiopoda anomopoda daphniidae daphnia ambigua scourfield, 1947 a north america u e 1, 7, 8, 12 daphnia parvula fordyce, 1901 a north america u e 7, 8 copepoda cyclopoida ergasilidae neoergasilus japonicus (harada, 1930) a asia u e 19 malacostraca isopoda asellidae proasellus banyulensis (racovitza, 1919) n.a. europe u e 3, 8 decapoda cambaridae procambarus clarkii (girard, 1852) a north america i e 5, 10, 17, 18 hexapoda insecta diptera culicidae aedes albopictus (skuse, 1894) a southeast asia u e 15 a, allochthonous; p, parautochthonous; i, intentional; u, unintentional; e, established; s, sporadic; n.a., not available; 1, calvo et al., 1993; 2, manganelli et al., 1995; 3, stoch et al., 1996; 4,weidema, 2000; 5, d’angelo and lo valvo, 2003; 6, zettler and richard, 2003; 7, marrone et al., 2005; 8, ruffo and stoch, 2006; 9, cianfanelli et al., 2007; 10, naselli-flores et al., 2007; 11, reitano et al., 2007; 12, gherardi et al., 2008; 13, liberto et al., 2010; 14, marrone et al., 2011; 15, carminade et al., 2012; 16, colomba et al., 2013; 17, di leo et al., 2014; 18, bellante et al., 2015; 19, alfonso and marrone, upublished data. no nco mm er cia l u se on ly 6alien species in sicilian inland water no specimens collected in sicilian inland waters were, to our knowledge, ever studied or described and whose presence itself in the island is to be considered dubious. the only aquatic bird ever reported as a nis for sicily is the mute swan, cygnus olor (gmelin, 1789) (scalera, 2001). however, the specimens overwintering in sicily are likely coming from the southern balkans, where the species is autochthonous as breeding as well as wintering bird, and should therefore be considered autochthonous in sicily (b. massa, personal communication). discussion the animal xenodiversity of sicilian inland waters when comparing the checklists reported in tabs. 1 to tab. 2. list of the non-indigenous vertebrate animal species known to occur in sicilian inland waters. taxon category origin introduction status source(s) allochthonous vs intentional vs established vs parautochthonous unintentional sporadic chordata osteichthyes perciformes centrarchidae micropterus salmoides lacépède, 1802 a north america i e 7, 12, 16, 17, 18, 21, 25 percidae perca fluviatilis linnaeus, 1758 n.a. eurasia i e 2, 18, 25 cypriniformes cyprinidae carassius auratus (linnaeus, 1758) n.a. asia i e 2, 3, 6, 7, 8, 16, 17, 18, 21, 25 cyprinus carpio (linnaeus, 1758) p* eurasia i e 2, 3, 6, 7, 16, 17, 18, 21, 22, 25, 27 rutilus rubilio bonaparte, 1837 a southern italy i e 4, 6, 7, 8, 12, 16, 17, 18, 28 tinca tinca (linnaeus, 1758) p europe i e 2, 3, 6, 7, 16, 17, 25 siluriformes ictaluridae ameiurus melas (rafinesque, 1820) a north america i e 6, 7, 12, 17, 18, 21, 25 cyprinodontiformes poecilidae gambusia holbrooki girard, 1859 a north america i e 1, 3, 6, 7, 8, 9, 15, 16, 17, 18, 21, 22, 25, 27, 29 esociformes esocidae esox cf. lucius§ a n.a. i e 6, 7, 12, 16, 17, 25 salmoniformes salmonidae oncorhynchus mykiss walbaum, 1792 a north america i s 6, 7, 16, 17, 18, 21 salmo trutta linnaeus, 1758 a europe i e 7, 16, 17, 18, 21, 25, 26 amphibia anura pipidae xenopus laevis (daudin, 1802) a africa i e 14, 19, 20, 21, 23, 24 reptilia testudines emydidae emys orbicularis s.l. a peninsular italy i n.a. 11, 30 trachemys scripta elegans (wied, 1839) a north america i s 19, 22 mammalia rodentia myocastoridae myocastor coypus molina, 1872 a south america i e 17, 19, 21 muridae rattus norvegicus berkenhout, 1769 p asia u e 10, 13, 17, 19, 21 a, allochthonous; p, parautochthonous; i, intentional; u, unintentional; e, established; s, sporadic; n.a., not available or not applicable; *its parautochthony considered dubious by gherardi et al. (2008, and references therein); §no sound information on the identity of the pikes introduced in sicily is available (see text); 1, consoli, 1928; 2, faranda et al., 1977; 3, tigano, 1983; 4, tigano and ferrito, 1986; 5, lo valvo et al., 1993; 6, ferrito and tigano, 1995; 7, tigano and ferrito, 1996; 8, russo et al., 1997; 9, veronesi et al., 1997; 10, sarà, 1998; 11, lenk et al., 1999; 12, russo et al., 1999; 13, scalera, 2001; 14, lillo et al., 2005; 15, duchi, 2006a; 16, duchi, 2006b; 17, ruffo and stoch, 2006; 18, nocita and zerunian, 2007; 19, aa.vv., 2008; 20, faraone et al., 2008; 21, gherardi et al., 2008; 22, termine et al., 2008; 23, lillo et al., 2011; 24, lillo et al., 2013; 25, bianco, 2014; 26, duchi, 2014a; 27, duchi, 2014b; 28, duchi, 2014c; 29, duchi and miceli, 2014; 30, vamberger et al., 2015. no nco mm er cia l u se on ly 7 f. marrone and l. naselli-flores 3 with the information currently available in literature (scalera, 2001; ruffo and stoch, 2006; nocita and zerunian, 2007; cianfanelli et al., 2007; gherardi et al., 2008; bianco, 2014), a certain decoupling is obvious, with the latter lacking some taxa which are actually present in sicily, or conversely including others whose presence is dubious or to be excluded for the island. this is to be ascribed to the shortage of studies explicitly focused at investigating sicilian xenodiversity, with the few existing data scattered among often difficult-to-get pieces of literature. furthermore, most of the whole-country-scale reviews are focusing on allochthonous taxa only, leaving aside the parautochthonous and translocated species, which are possibly difficult to single out when working on large geographical scales, but which can be easily identified when the study area is of limited extension and geographically well-defined as the present case-study. the case of bony fishes is emblematic: although they tab. 3. checklist of the taxa whose presence in sicily and\or whose non-indigenous status is uncertain. taxon origin notes source(s) mollusca gastropoda lymnaeidae galba truncatula (o.f. müller, 1774) europe the species is considered autochthonous by liberto et al., 2010 5 planorbidae helisoma anceps (menke, 1830) north america it might have been confused with the congener h. duryi 5 (cf. cianfanelli et al., 2007) bivalvia corbiculidae corbicula fluminea (o.f. müller, 1774) southeast asia the report is based on a single specimen stored in the mollusc collection 2 of the hebrew university of jerusalem crustacea branchiopoda anomopoda moinidae moina affinis birge, 1893 north america possibly a misidentification for the congener species moina salina 1 or m. brachiata (cf. marrone et al., 2005) chordata osteichthyes cypriniformes cyprinidae carassius carassius linnaeus, 1758 europe taxon reported for sicily without providing precise locality data 6 nor relevant references squalius cf. cephalus § n.a. taxon reported for sicily without providing precise locality data 4 nor relevant references pseudorasbora parva asia taxon reported for sicily without providing precise locality data 7, 8 (temminck and schlegel, 1846) nor relevant references perciformes centrarchidae lepomis gibbosus (linnaeus, 1758) north america taxon reported for sicily without providing precise locality data 7, 8, 9 nor relevant references siluriformes ictaluridae ameiurus nebulosus (lesueur, 1819) north america taxon reported for sicily without providing precise locality data 7, 8, 9 nor relevant references reptilia testudines geoemydidae mauremys cf. sinensis (gray, 1834) asia no evidences on the occurrence of the species in the wild are available 10 aves anseriformes anatidae cygnus olor (gmelin, 1789) eurasia the individuals overwintering in sicily might come from the balkans, 3, 4 where the species is autochthonous. they should thus not be considered nis in sicily (b. massa, pers. comm.) §no sound information on the species-level identity of the taxon\taxa possibly introduced to sicily is available; n.a., not available or not applicable; 1, faranda, 1977; 2, mienis, 1991; 3, lo valvo et al., 1993; 4, scalera, 2001; 5, zettler and richard, 2003; 6, duchi, 2006b; 7, nocita and zerunian, 2007; 8,gherardi et al., 2008; 9, bianco, 2014; 10, panzeri et al., 2014. no nco mm er cia l u se on ly 8alien species in sicilian inland water are by far the most represented non-indigenous taxon occurring in sicily, only south-eastern sicilian inland water bodies have been investigated (ferrito and tigano, 1995; duchi, 2006b, and references therein) to date, and nearly no information is available for most of sicilian inland waters (but see russo et al., 1999; duchi, 2014c). as a consequence, no updated local reviews are available, and the italian lists are rather incomplete or include some taxa whose actual presence in the island is dubious or can be excluded (ruffo and stoch, 2006; nocita and zerunian, 2007; gherardi et al., 2008; tricarico et al., 2010; and, partly, bianco, 2014). furthermore, five out of the 11 fish species reported in tab. 2 are likely translocated taxa from peninsular italy, where they are autochthonous, and thus partly or completely overlooked by nocita and zerunian (2007), gherardi et al. (2008), and bianco (2014). when dealing with biological invasions, the translocation of fish and other animals among different freshwater ecosystems within the region should also be considered. quite often we actually observed fish introductions in temporary ponds where they can significantly alter and threaten the structure of the native biota (naselli flores and barone, 2012). the strong predominance of vertebrate over invertebrate taxa in tabs. 1-3 is likely an artefact due to the high visibility of the former, and to the inadequate number of taxonomists present in sicily for several aquatic invertebrate taxa. it is quite evident that the presence of a single non-indigenous insect species (i.e., the asian tiger mosquito, aedes albopictus) in the list here presented has to be ascribed to the paucity of information currently available on sicilian inland water insect communities rather than to the actual presence of a single nis belonging to this important and species-rich taxon. more accurate information is available for inland water molluscs and crustaceans, two among the best-studied invertebrate taxa in sicily; conversely, no information is available for other important taxa as porifera, cnidaria, or annelida, which include several species known to be introduced in italian and european inland waters and that might well be present in sicilian inland waters as well. in spite of the patent incompleteness of the currently available nis checklist, seven out of the 31 nis positively present in sicilian inland waters are listed among the 100 of the world’s worst invasive alien species (lowe et al., 2000), i.e. one mollusc (the zebra mussel, d. polymorpha), one insect (the asian tiger mosquito, a. albopictus), three fish (the brown trout s. trutta, the carp c. carpio, the large-mouth bass m. salmoides), one reptile (the redeared slider, t. scripta elegans), and one mammal (the coypu, m. coypus). at least two other heavily invasive species should be added to this list, i.e. the eastern mosquito-fish, gambusia holbrooki, and the african clawed frog, xenopus laevis, whose impact on native biotas is largely known and also verified for sicilian inland waters (lillo et al., 2011; duchi and micieli, 2014). overall, at least nine highly invasive species threatening the native inland water biota are present in sicily. unfortunately, only few studies have been to date addressed to the evaluation of the impact of nis on sicilian native species and ecosystems, and these are mostly dealing with fish (ferrito and tigano, 1996; russo et al., 1999; duchi, 2006a; duchi and micieli, 2014; duchi et fig. 1. histogram of the nis in sicilian inland waters by phylum. among brackets the number of confirmed species in the island. no nco mm er cia l u se on ly 9 f. marrone and l. naselli-flores al., 2014a); however, some information on the potential negative role of the red swamp crayfish (p. clarkii) in sicilian ecosystems as a vector for toxins and heavy metals is available (naselli-flores et al., 2007; bellante et al., 2015), as well as sound evidences on the threats exerted by the african clawed frogs on native amphibians (faraone et al., 2008; lillo et al., 2011) (tab. 4). is sicily a hot-spot for inland water xenodiversity? gherardi et al. (2008) and tricarico et al. (2010), based on a dataset including only those species which are allochthonous at the whole-country-level (i.e., without considering the parautochthonous and the translocated species), pointed out that northern italy is the hot-spot of italian inland water xenodiversity. conversely, boggero et al. (2014), based on a different dataset aimed at exploring the susceptibility to invasions of different habitats and regions, found out that the number of alien species (considering only the allochthonous taxa, according to the definitions used in this paper) is a correlate of temperature, so that sites in warmer areas host in average more alien species than those located in colder ones, which is likely due to a more intense touristic frequentation of the former. the apparent contrast among these results is due to the lacking or inadequacy of sampling surveys and published data for the southern regions of the country: when the mere number of different nis known to occur in different regions is compared, the best-studied areas (i.e., the central and northern italian ones) show the higher number of nis, which is in fact a function of the sampling and publishing efforts rather than a faithful mirror of the actual xenodiversity distribution pattern. conversely, when a balanced subset of soundly comparable study sites is infig. 2. histogram of the biogeographical regions of origin for the nis confirmed in sicilian inland waters. among brackets the number of species originating from any biogeographical region. melanoides tuberculata native range lies in both afrotropical and palaearctic regions, and is here reported under the bar named other distribution patterns. tab. 4. published data on the impact of nis on the sicilian autochthonous inland water biota. taxon recorded impact on the indigenous biota source procambarus clarkii vector of toxins and heavy metals to higher trophic levels naselli-flores et al., 2007; bellante et al., 2015 micropterus salmoides predation on salaria fluviatilis russo et al.. 1999 gambusia holbrooki competition with aphanius fasciatus duchi, 2006a; duchi and micieli, 2014 salmo trutta hybridization with salmo cettii duchi, 2014a xenopus laevis predation on indigenous amphibians lillo et al., 2011 no nco mm er cia l u se on ly 10alien species in sicilian inland water vestigated, the result is exactly the opposite, with the warmer southern italian regions being more prone to be successfully invaded than the central and northern ones. such a perspective makes even more urgent the need of properly investigating southern italian and sicilian inland water biota, possibly an actual but overlooked hot-spot of inland water animal xenodiversity in europe. acknowledgments bruno massa (palermo) is gratefully acknowledged for the comments he provided on the autochthonous status of the mute swan (cygnus olor) specimens wintering in sicily. two anonymous reviewers are acknowledged for their comments, which helped to improve a first draft of the manuscript. references aa.vv, 2008. atlante della biodiversità della sicilia: vertebrati terrestri. collana “studi e ricerche” 6, arpa sicilia, palermo, 536 pp. airoldi l, turon x, perkol-finkel s, rius m, 2015. corridors for aliens but not for natives: effects of marine urban sprawl at a regional scale. diversity distrib. 21:755-768. bellante a, maccarone v, buscaino g, buffa g, filiciotto f, traina a, del core m, mazzola s, sprovieri m, 2015. trace element concentrations in red swamp crayfish (procambarus clarkii) and surface sediments in lake preola and gorghi tondi natural reserve, sw sicily. environ. monit. assess. 187: 404. bianco pg, 2014. an update on the status of native and exotic freshwater fishes of italy. j. appl. ichthyol. 30:62-77. bianco pg, delmastro gb, 2011. recenti novità tassonomiche riguardanti i pesci d’acqua dolce autoctoni in italia e descrizione di una nuova specie di luccio. researches on wildlife conservation 2:(suppl.)1-13. blanchet s, 2012. the use of molecular tools in invasion biology: an emphasis on freshwater ecosystems. fish. manag. ecol. 19:120-132. boggero a, basset a, austoni m, barbone e, bartolozzi l, bertani i, campanaro a, cattaneo a, cianferoni f, corriero g, dŏrr am, elia ac, ficetola gf, kamburska l, la porta g, lauceri s, ludovisi a, gaino e, goretti e, lorenzoni m, manca m, marchetto a, morabito g, nonnis marzano f, oggioni a, pierri c, riccardi n, rossetti g, ungaro n, volta p, zaupa s, fontaneto d, 2014. weak effects of habitat type on susceptibility to invasive freshwater species: an italian case study. aquat. conserv. 24:841-852. calvo s, barone r, naselli-flores l, fradàorestano c, dongarrà g, lugaro a, genchi g, 1993. limnological studies on lakes and reservoirs of sicily. naturalista siciliano 27(suppl.):1-292. carminade c, medlock jm, ducheyne e, mcintyre km, leach s, baylis m, morse ap, 2012. suitability of european climate for the asian tiger mosquito aedes albopictus: recent trends and future scenarios. j. r. soc. interface 9:2708-2717. chandra s, gerhardt a, 2008. invasive species in aquatic ecosystems: issue of global concern. aquat. invasions 3:1-2. chytrý m, jarošík v, pyšek p, hájek o, knollová i, tichý l, danihelka j, 2008. separating habitat invasibility by alien plants from the actual level of invasion. ecology 89:1541-1553. cianfanelli s, lori e, bodon m, 2007. non-indigenous freshwater molluscs and their distribution in italy, p. 103-121. in: f. gherardi (ed.), biological invaders in inland waters: profiles, distribution, and threats. springer, dordrecht. colomba ms, liberto f, reitano a, grasso r, di franco d, sparacio i, 2013. on the presence of dreissena polymorpha pallas, 1771 and sinanodonta woodiana woodiana (lea, 1834) in sicily (bivalvia). biodivers. j. 4:571-580. consoli, n, 1928. gl’interventi di piccola bonifica nella lotta contro la malaria in sicilia. rivista sanitaria siciliana 18, francesco sanzo e c., industria tipografica editrice, palermo: 108 pp. cox jg, lima sl, 2006. naiveté and an aquatic – terrestrial dichotomy in the effects of introduced predators. trends ecol. evol. 21:674-680. d’angelo s, lo valvo m, 2003. on the presence of the red swamp crayfish procambarus clarkii (girard, 1852) (decapoda cambaridae) in sicily (italy). naturalista siciliano 27:325-327. di leo c, faraone fp, lo valvo m, 2014. a new record of the red swamp crayfish, procambarus clarkii (girard, 1852) (crustacea, cambaridae), in sicily, italy. biodivers. j. 5:425-428. duchi a, 2006a. osservazioni sui popolamenti di nono (aphanius fasciatus, valenciennes) e gambusia (gambusia holbrooki, girard) in provincia di ragusa. biologia ambientale 20:73-75. duchi a, 2006b. distribuzione della fauna ittica nelle acque interne dell’areale ibleo: la provincia di ragusa. biologia ambientale 20:291-294. duchi a, 2014a. relazione tra lunghezza e frequenza fenotipica di salmo cettii/salmo trutta ed i loro ibridi in un corso d’acqua siciliano (irminio, ragusa). ital. j. freshwater ichthyol. 1:207-210. duchi a, 2014b. indagini sulla fauna ittica dei corsi d’acqua presenti in alcuni siti di importanza comunitaria della piana di gela (cl, sicilia, italia). ital. j. freshwater ichthyol. 1:211214. duchi a, 2014c. ampliamento dell’areale della rovella (rutilus rubilio, bonaparte, 1837) in sicilia: nuove segnalazioni nell’areale ibleo ed in provincia di trapani. ital. j. freshwater ichthyol. 1:221-224. duchi a, micieli g, 2014. prima segnalazione di gambusia gambusia holbrooki girard, 1859 (cyprinodontiformes poecilidae) nel pantano longarini: un fattore di rischio per il popolamento di nono aphanius fasciatus valenciennes, 1821 (cyprinodontiformes cyprinodontidae). naturalista siciliano 38:115-117. ehrenfeld jg, 2010. ecosystem consequences of biological invasions. annu. rev. ecol. evol. syst. 41:59-80. elton cs, 1958. the ecology of invasions by animals and plants. university of chicago press, chicago and london (ed. june 2000): 196 pp. faranda f, 1977. primo censimento delle aree destinabili ad acquacoltura in sicilia. atti società peloritana scienze fisiche, matematiche e naturali 23(suppl.):1-113 no nco mm er cia l u se on ly 11 f. marrone and l. naselli-flores faraone fp, lillo f, giacalone g, lo valvo m, 2008. the large invasive population of xenopus laevis in sicily (italy). amphibia-reptilia 29:405-412. ferrito v, tigano c, 1995. the distribution of the ichthyofauna in the simeto basin (sicily). cybium 19:187-198. ferrito v, tigano c, 1996. decline of aphanius fasciatus (cyprinodontidae) and salaria fluviatilis (blenniidae) populations in freshwaters of eastern sicily. ichthyol. explor. freshwaters 7:181-184. ficetola f, thuiller w, padoa-schioppa e, 2009. from introduction to the establishment of alien species: bioclimatic differences between presence and reproduction locality in the slider turtle. diversity distrib. 15:108-116. gazzetta ufficiale, 2015. elenco delle specie alloctone escluse dalle previsioni dell’articolo 2, comma 2 -bis , della legge n. 157/1992. decreto 19 gennaio 2015. gazzetta ufficiale della repubblica italiana, serie 31, 07/02/2015, pp 5-6. accessed on: 27 august 2015. available from: http://www.minambiente.it/sites/default/files/archivio/allegati/biodiversita/ dim_07_02_2015_specie_alloctone.pdf gherardi f, 2006.bioinvasions in fresh waters and the nerodilemma. pol. j. ecol. 4:549-561. gherardi f, 2007. biological invaders in inland waters: profiles, distribution, and threats. springer, dordrecht: 733 pp. gherardi f, bertolino s, bodon m, casellato s, cianfanelli s, ferraguti m, lori e, mura g, nocita a, riccardi n, rossetti g, rota e, scalera r, zerunian s, tricarico e, 2008. animal xenodiversity in italian inland waters: distribution, modes of arrival, and pathways. biol. invasions 10:435-454. iucn, 2000. iucn guidelines for the prevention of biodiversity loss caused by alien invasive species. available from: https://portals.iucn.org/library/efiles/documents/rep-2000052.pdf keller rp, geist j, jeschke jm, kühn i, 2011. invasive species in europe: ecology, status, and policy. environ. sci. eur. 23:23. lenk p, fritz u, joger u, wink m, 1999. mitochondrial phylogeography of the european pond turtle, emys orbicularis (linnaeus 1758). mol. ecol. 8:1911-1922. leppäkoski e, gollasch s, olenin s, 2002. alien species in european waters, p. 1-6. in: e. leppäkoski, s. gollasch and s. olenin (eds.), invasive aquatic species of europe: distribution, impact and management. kluwer academic, dordrecht. liberto f, giglio s, reitano a, colomba ms, sparacio i, 2010. molluschi terrestri e dulciacquicoli di sicilia della collezione f. minà palumbo di castelbuono. monografie naturalistiche, 2. edizioni danaus, palermo: 136 pp. lillo f, marrone f, sicilia a, castelli g, zava b, 2005. an invasive population of xenopus laevis (daudin, 1802) in italy. herpetozoa 18:63-64. lillo f, faraone fp, lo valvo m, 2011. can the introduction of xenopus laevis affect native amphibian populations? reduction of reproductive occurrence in presence of the invasive species. biol. invasions 13:1533-1541. lillo f, dufresnes c, faraone fp, lo valvo m, stöck m, 2013. identification and potential origin of invasive clawed frogs xenopus (anura: pipidae) in sicily based on mitochondrial and nuclear dna. ital. j. zool. 80:566-573. lo valvo m, massa b, sarà m, 1993. uccelli e paesaggio in sicilia alle soglie del terzo millennio. naturalista siciliano 17 (suppl.):1-371. lowe s, browne m, boudjelas s, de poorter m, 2000. 100 of the world’s worst invasive alien species. a selection from the global invasive species database. aliens 12:1-12. manganelli g, bodon m, favilli l, giusti f, 1995.gastropoda pulmonata p. 1-60. in: a. minelli, s. ruffo and s. la posta (eds.), checklist delle specie della fauna italiana, 16. edizioni calderini, bologna. marr sm, olden jd, leprieur f, arismendi i, ćaleta m, morgan dl, nocita a, šanda r, tarkan as, garcía-berthou e, 2013. a global assessment of freshwater fish introductions in mediterranean-climate regions. hydrobiologia 719:317-329. marrone f, barone r, naselli-flores l, 2005. cladocera (branchiopoda: anomopoda, ctenopoda, and onychopoda) from sicilian inland waters: an updated review. crustaceana 78:1025-1039. marrone f, lo brutto s, arculeo m, 2011. cryptic invasion of a gastropod mollusc in southern europe: the case of ferrissia fragilis (pulmonata: ancylidae). biologia 66:484-490. mazza g, aquiloni l, inghilesi af, giuliani c, lazzaro l, ferretti g, lastrucci l, foggi b, tricarico e, 2015. aliens just a click away: the online aquarium trade in italy. management of biological invasions 6:253-261. mienis hk, 1991. some remarks concerning asiatic clams invading europe with a note on sample of corbicula fluminea (mueller, 1774) from trapani, sicily. notiziario sim 9:137139. naiman rj, dudgeon d, 2011. global alteration of freshwaters: influences on human and environmental well-being. ecol. res. 26:865-873. naselli-flores l, barone r, 2012. phytoplankton dynamics in permanent and temporary mediterranean waters: is the game hard to play because of hydrological disturbance? hydrobiologia 698:147-159. naselli-flores l, barone r, marrone f, d’angelo s, 2007. 100 milioni di microcystis spp. + 5 procambarus clarkii = 0 emys trinacris, ovvero tossine, invasori ed estinzione nei gorghi tondi, laghi salmastri della sicilia sud-occidentale, p. 76-77. proceedings joined meet. aiol-site, ancona, italy. nocita a, zerunian s, 2007. l’ittiofauna aliena nei fiumi e nei laghi d’italia. biologia ambientale, 21: 93-96. panzeri m, mori e, mazza g, menchetti m, 2014. records of introduced stripe-necked terrapins (mauremys species) in italy. acta herpetologica 9:227-230. pfenninger m, schwenk k, 2007. cryptic animal species are homogeneously distributed among taxa and biogeographical regions. bmc evol. biol. 7:121. pimentel d, zuniga r, morrison d, 2005. update on the environmental and economic costs associated with alien-invasive species in the united states. ecol. econ. 52:273-288. reitano a, liberto f, sparacio i, 2007. nuovi dati su molluschi terrestri e dulciacquicoli di sicilia. 1° contributo (gastropoda prosobranchia neotaenioglossa; gastropoda pulmonata basommatophora, stylommatophora). naturalista siciliano 31:311-330. ruffo s, stoch f, 2006. checklist and distribution of the italian fauna. 10,000 terrestrial and inland waters species. memorie del museo civico di storia naturale di verona, serie 2, sezione scienze della vita 17:1-307. russo g, violani c, zava b, 1997. observations on population dynamics of rutilus rubilio cyprinidae in a man-made hyno nco mm er cia l u se on ly 12alien species in sicilian inland water pertrophic basin (arancio lake, southwest sicily). ital. j. zool. 65(suppl.):549-551. russo g, la rocca s, violani c, zava b, 1999. contributions to the knowledge of sicilian freshwater fishes. ii. notes on some allochthonous species recently introduced. doriana 7:1-7. sarà m, 1998. i mammiferi delle isole del mediterraneo. l’epos società editrice, palermo,166 pp. saltonstall k, 2002. cryptic invasion by a non-native genotype of the common reed, phragmites australis, into north america. p. natl. acad. sci. usa 99:42445-42449. scalera r, 2001. invasioni biologiche. le introduzioni di vertebrati in italia: un problema tra conservazione e globalizzazione. collana verde, 103. corpo forestale dello stato. ministero delle politiche agricole e forestali, roma: 368 pp.. simberloff d, 2003. confronting introduced species: a form of xenophobia? biol. invasions 5:179-192. simberloff d, martin jl, genovesi p, maris v, wardle da, aronson j, courchamp f, galil b, garcía-berthou e, pascal m, pyšek p, sousa r, tabacchi e, vilà m, 2013. impacts of biological invasions: what’s what and the way forward. trends ecol. evol. 28:58-66. sowerby gb, 1873-1874. monograph of the genus physa, pp. 27. in: reeve, lovell, 1843-78, conchologica iconica. vol. 19. reeve, benham and reeve, london. stoch f, valentino f, volpi e, 1996. taxonomic and biogeographic analysis of the proasellus coxalis-group (crustacea, isopoda, asellidae) in sicily, with description of proasellus montalentii n.sp. hydrobiologia 317:247-258. termine r, canale ed, ientile r, cuti n, di grande cs, massa b, 2008. vertebrati della riserva naturale speciale e sito di importanza comunitaria lago di pergusa. naturalista siciliano 32:105-186. tigano c, 1983. dati preliminari sul popolamento ittico del fiume simeto (sicilia). boll. acc. gioenia sci. nat. 16:81-98. tigano c, ferrito v, 1986. sulla presenza di rutilus rubilio (bp. 1837) in sicilia (pisces, cyprinidae). animalia 13:109-124. van bocxlaer b, clewing c, mongindo etimosundja jp, kankonda a, ndeo ow, albrecht c, 2015. recurrent camouflaged invasions and dispersal of an asian freshwater gastropod in tropical africa. bmc evol. biol. 15:33. veronesi r, bellini r, celli g, 1997. ruolo di gambusia holbrooki nel contenimento dei culicidi e suo impatto sulle biocenosi acquatiche. biologia ambientale 3:24-40. vamberger m, stuckas h, sacco f, d’angelo s, arculeo m, cheylan m, corti c, lo valvo m, marrone f, wink m, fritz u, 2015. differences in gene flow in a twofold secondary contact zone of pond turtles in southern italy (testudines: emydidae: emys orbicularis galloitalica, e. o. hellenica, e. trinacris). zool. scripta 44:233-249. weidema ir, 2000. introduced species in the nordic countries. nord 13:1-242. williamson m, 1996. biological invasions. chapman and hall, london: 224 pp. zettler ml, richard d, 2003. kurze bemerkungen über süsswassermollusken siziliens unterbesonderer berücksichtigung von theodoxus meridionalis (philippi, 1836). malak. abh. 21:29.38. no nco mm er cia l u se on ly layout 1 introduction about 50,000 reservoirs with dams higher than 15 m exist in the world and many huge dam construction projects are in the planning phase (wcd, 2000). especially in the tropics and sub-tropics, the number of recently constructed reservoirs is high. the area covered by those reservoirs is increasing and already makes up an area of about 500,000 km², which equals one third of the surface area of non-artificial surface water bodies (wcd, 2000). by impounding rivers through the construction of dams, riverine systems and biochemical cycles are disrupted (friedl and wüest, 2002). suspended solids and transported material are trapped and settle in the reservoirs (odhiambo and boss, 2004). the sediment quality and amount of sediment have a direct influence on the management of the reservoir as well as on water quality in the reservoir. it is therefore necessary to characterise and manage the accumulated sediment volumes. as the spatial dimensions of many of the reservoirs are huge, it is still a problem to obtain representative data for several quality parameters of the sediment. echo sounding systems have been found to allow for a timeand cost-saving data acquisition, especially when they are used on larger spatial scales (freitas et al., 2006; anderson et al., 2008; poulain et al., 2011). they represent the most promising approach to extensive sediment classification apart from traditional point sampling techniques like grab sampling and sediment coring. multibeam echo sounders (mbs) and swath systems provide the highest coverage of area in relation to needed vessel time, but are inherently more expensive than single beam echo sounders (sbs). in regards of seabed classification literature shows successful studies for mbs (bentrem et al., 2002; preston, 2009; hamilton and parnum, 2011) and sbs (tęgowski, 2005; anderson and pacheco, 2011; poulain et al., 2011), both systems have advantages and disadvantages. mbs and sbs provide high accuracy during seabed classification. however, the sbs need dense line spacing for comparable spatial results. the sbs features the advantage that the second bottom echo can also be integrated in the seabed classification (parnum et al., 2009). systems like the simrad ea 400 use two frequencies at the same time, which may improve the sediment classification due to a lower second frequency, which allows for a better classification of the volume characteristics due to deeper sediment penetration. independent from the selected echo sounder, the classification of sediment parameters is still largely limited to physical parameters, such as grain size distribution. few readvances in oceanography and limnology, 2016; 7(1): 93-105 article doi: 10.4081/aiol.2016.5623 investigation of echo sounding parameters for the characterisation of bottom sediments in a sub-tropical reservoir stephan hilgert,* adrian wagner, lisa kiemle, stephan fuchs institute for water and river basin management (iwg), division of aquatic environmental engineering (isww), karlsruhe institute of technology (kit), gotthard-franz-str.3, bld. 50.31, 76131 karlsruhe, germany *corresponding author: stephan.hilgert@kit.edu abstract the increasing number of reservoirs around the world today reaches a surface area of around 500,000 km², equalling one third of that of non-artificial surface water bodies. by impounding rivers through the construction of dams, riverine systems and biochemical cycles are disrupted. different types of transported materials are trapped behind the dams and form layers of sediment. a comprehensive combination of two frequencies with four pulse lengths were tested in order to classify multiple physical and chemical sediment parameters in the vossoroca reservoir in the southeast of brazil, paraná state. a number of core and grab samples was taken and analysed for a variety of chemical and physical parameters. these data served as ground truthing for the hydro-acoustic assessment of the sediment. eight hydro-acoustic parameters were derived from the echo signals obtained with an ea 400 system using the sonar5-pro software. the major objective of defining the optimal survey parameters for the echo sounder as well as determining the difference between core and grab samples was reached by correlating the various single parameters and identifying the best combinations. density and grain size distribution represented the best detectable sediment features with r-values of 0.94 and 0.95. the lower 38 khz frequency generally had a better performance than the 200 khz frequency. results show that core samples reached a significantly higher quality of correlation for sediment characterisation. additionally, it was found that shorter pulse lengths yield a better characterisation. the results underline the potential of single beam echo sounders for extensive sediment characterisation. this methodology may be used for future mass balance estimations of large reservoirs. key words: echo sounding; sediment characterisation; pulse length; sediment sampling; sbc, reservoir. received: november 2015. accepted: june 2016. no nco mm er cia l u se on ly 94 s. hilgert et al. sults have been published so far with respect to correlations between other sediment parameters, e.g. organic content, iron content, phosphate content, and loss on ignition, and hydro-acoustic parameters. next to the influence of the named sediment parameters, this study investigates different pulse length in order to find optimal configurations for seabed classification. apart from the best echo sounder configuration for the detection of the parameters listed, the influence of the sediment sampling method was investigated. since grab and core sampling produce different sediment sample types, a significant influence on the ground truthing results can be expected. most of the data published refer to the correlation of hydro-acoustic information with grab samples taken by e.g. an ekman dredge (bentrem et al., 2002; amiri-simkooei et al., 2011; anderson and pacheco, 2011; poulain et al., 2011; anderson and martinez, 2015). several well-established commercial echo sounding systems and related software for lakebed classification are available on the market (i.e., qtc view, etc.), internal data processing, however, is a black box to the user. we used a combination of a kongsberg ea 400 single beam echo sounder with frequencies of 200 and 38 khz and the sediment classification tool of the sonar5-pro post-processing software for lakebed characterization. the formulas for the seabed classification used in the sonar5-pro software are openly documented and intermediary results can be exported and processed e.g. in a combination with matlab (mathworks®) (balk et al., 2011). the data were collected in march and november 2011 during two surveys of the vossoroca reservoir in the southeast of brazil. during the first campaign, bathymetry of the entire reservoir was obtained in a rather dense grid. the maximum depth of the reservoir was found to reach 17 metres. subsequently, a 3d model was created, which served to plan the following survey in november. eight different hydro-acoustic parameters were derived from the echo signals for later correlation with the sediment properties at the corresponding locations. the sediment properties included physical parameters, such as grain size distribution, but also chemical parameters, such as phosphorus and iron contents. in addition to the standardised ‘first echo division method’ and ‘first/second bottom ratio method’ (burczynski, 1999), further signal parameters extracted by the sediment classification tool of the sonar5pro software were tested with respect to their correlations with sediment parameters obtained by ground truthing and at changed pulse durations. this study was aimed at finding the optimal echo sounder configuration for the characterisation of sediments using a non-commercial software. additionally, the influence of the sediment sampling method was analysed. methods study area the vossoroca reservoir is located in southern brazil in the state of paraná, approximately 50 km southeast of curitiba in the ‘serra do mar’ mountain range at approximately 833 m asl. the annual rainfall is about 1900 mm, the climate is subtropical. the reservoir ,which covers an area of about 5 km² and has a capacity of 33.6 x 106 m³ (republic of brazil, 1969), was created in 1940 to control water flow for the chaminé hydroelectric power plant located 7 km downstream. the average water depth is about 8 m, maximum depth is about 17 m. the catchment area of the reservoir has a size of 151 km² (republic of brazil, 1969) and is predominantly rural. the immediate surroundings of the reservoir are used for recreation, whereas the upstream area of the basin is mainly used by agriculture. the major part of the reservoir belongs to the nature protection area of guaratuba. bathymetric survey covered the entire lake except for the very shallow parts (less than 1.30 m). in these areas, no proper echo signal could be received and the risk of running aground was too high. fig. 1 shows the bathymetric measurement grid. point sampling focused on the left arm and the central part of the reservoir to facilitate the sampling process (fig. 2). hydro-acoustic data acquisition a kongsberg ea 400 single beam echo sounder with frequencies of 200 and 38 khz was used for the acquisition of hydro-acoustic data. tab. 1 reports the echo sounder characteristics. the ea 400 features opening angles of 7° for 200 khz and longitudinal and transverse angles of 13° and 21°, respectively, at 38 khz. the survey was carried out to characterise the lakebed under very shallow conditions (2-17 m depth). the transducer was mounted vertically (0°angle) on the starboard side of a small aluminium vessel, 67 cm below the water surface to prevent surface bubbles from interfering with the meastab. 1. overview of the ea 400 echo sounder characteristics. frequency opening angle (longitudinal) opening angle (transverse) pulse length (ms) power input (w) 200 khz 7° 7° 0.064; 0.128; 0.256; 0.512 100-1000 38 khz 13° 21° 0.256; 0.512; 1.024; 2.058 100-1000 no nco mm er cia l u se on ly hydro-acoustic sediment characterisation 95 urements. the average vessel speed was about 2 m s–1. due to the calm weather and, hence, the absence of waves, the influence of pitch, roll, and heave could be neglected. every day before the acoustic measurements, the actual water level of the reservoir was determined with the help of a reference pole. in this way, the acoustic signals could be referenced to the same water level and merged and compared. in addition, a temperature and a conductivity profile were recorded daily and both data were included in signal correction during post-processing. the temperature and conductivity profiles were used to correct the sound speed in the water phase during postprocessing. a leica gps 1200+ was used during all measurements. a rtk base station was positioned on a close by hill to allow for high precision of the positioning, below 1 m accuracy. the gps was connected to the notebook and transmitted the position signal to the ea 400 software. the ground at each position was sounded for the duration of at least 300 pings using the ea 400 echo sounder. as the survey focused on the potential of the echo sounder to obtain sediment characteristics, variable pulse length combinations at both frequencies were used, thus resulting in a data set of 300 pings per configuration and four configurations at each of the 32 positions. eight different hydro acoustic parameters were derived from the echo signals for later correlation with the sediment properties at the corresponding locations. bathymetric survey the bathymetric survey of the entire reservoir was made along a pre-defined survey grid. the transverse lines of the general survey grid had a track distance of approximately 50 to 100 m, while the western arm of the reservoir was surveyed at higher resolution. higher numbers of transverse lines were chosen to allow for better detecfig. 1. driven raster of the bathymetric survey in the vossoroca reservoir; the grey lines represent the driven tracks of the boat. no nco mm er cia l u se on ly 96 s. hilgert et al. tion of steep slopes. in addition to the longitudinal and transverse lines, a route close to the reservoir banks was driven to obtain a more complete data set of the shallow parts of the reservoir. to drive the boat along the pre-defined grid lines, a compass was installed on the boat for undelayed course correction. while measuring, the boat was kept as calm as possible, fast acceleration and swinging were prevented in order to reduce distorted signal detection. small changes in in heave, role or pitch were neglected during the bathymetric study. the measurements were carried out at both frequencies with a pulse length of 0.256 ms and a ping rate of 20 pings per second. the power input was set to 100 w to prevent non-harmonic distortions at 200 khz. sediment characterisation depth distribution of the digital lake model was used as a basis for the positioning of the sediment sampling points. the positions were selected in transect lines perpendicular to the general geomorphological shape and the former river bed (fig. 2). in case of very steep slopes or other inaccessible locations, the sample sites were shifted from the transects to the next possible position. during the second survey, 11 sediment cores and 21 grab samples were collected at 32 different positions. bathymetric information was used to select sampling positions, including not only the depth spectrum, but also site characteristics. to increase the density of the sampling points, the left arm and the central part of the reservoir were selected for detailed investigation. all sampling points were located in this part of the reservoir. the vessel was anchored with three anchors to ensure a stable position during the sampling and echo sounding phase. no movement of the boat occurred during data acquisition. ensonification of the sediment with the echo sounder was carried out before disturbing the sediment by taking core or grab samples. four different configurations were used fig. 2. positions of the core and grab sampling sites in the vossoroca reservoir. no nco mm er cia l u se on ly hydro-acoustic sediment characterisation 97 in order to find the best setting for sediment characterisation (tab. 2). all configurations were set to an input power of 100 w. as the variability of the signal between two successive pings may be high (burczynski, 1999), the echo sounder was set to all configurations one after another and each sampling position was ensonified for the duration of approximately one minute with 5 pings per second, resulting in a minimum of 300 pings per sampling point and configuration. bottom detection was carried out using the sonar5-pro bottom detection tool (balk et al., 2011). as a threshold for bottom detection, -36 db was chosen, as it showed best results for the automatic detection of the sediment water interface. each detected bottom line was controlled manually on pixel level to ensure correctness of bottom detection (e.g., to exclude trunks). post-processing of the echo sounder data resulted in a set of eight parameters for each ping (fig. 3). afterwards, all erroneous and extreme values were sorted out and mean values of the 300 pings were calculated for all eight parameters. this processing step was conducted in matlab r2014a (mathworks®). according to the different phases of bottom ensonification, the signal of an echo is divided into two parts (e1’ and e1) as shown in fig. 3 (burczynski, 1999). the first part covers the attack phase (duration: one pulse length from the bottom detection point) and the second the decay phase (duration: three pulse lengths from the end of the attack phase). as that part of the echo, which is created during the attack phase, is mainly caused by the bottom surface, the energy of this part of the echo (e1’) can be used as a measure of acoustic hardness or reflectivity. the part of the echo created during the decay phase is caused by diffuse backscattering from the sediment volume. as scattering depends on bottom roughness, the energy of this second part of the echo (e1) is generally described as acoustic roughness (burczynski, 1999). fig. 3. division of the first (left side) and second (right side) bottom echo into six basic hydro-acoustic parameters. tab. 2. overview of the four ea 400 echo sounder configurations with different pulse lengths. frequency (khz) configuration a configuration b configuration c configuration d pulse length (ms) pulse length (ms) pulse length (ms) pulse length (ms) 38 0.256 0.256 0.256 0.512 200 0.064 0.128 0.256 0.512 no nco mm er cia l u se on ly 98 s. hilgert et al. using the sonar 5 seabed classification tool, the attack value (hardness/ attsv1/ e1’) and decay values (roughness/ decsv1/ e1) were exported for the first bottom echo as well as for the second bottom echo, respectively the first multiple reflection (attsv2 and decsv2) (fig. 3). attsv2 and decsv2 sum up to the e2 value for the entire second bottom echo. attsv1 (e1’) and decsv1 (e1) describe the average volume backscattering strength sv1 [db] during the attack and decay phases of the first echo, respectively (orlowski, 1984). to calculate the average volume backscattering strength during either attack or decay phase, the echo strengths (sv1i) of each single sample belonging to that phase are converted into intensities, summarised, divided by the number of samples, and converted back into a db value (equations 1 and 2) (balk et al., 2011). this was performed for the first as well as for the second bottom echo. (eq.1) (eq. 2) in addition to the ‘basic’ attack and decay values, two values derived from the ‘first echo division method’ (e1’/e1) and the ‘first/second bottom ratio method’ (e1/e2) (orlowski, 1984) were exported. as further parameters, one value representing the reflected energy of the entire first bottom echo ‘attdecsv1’ (e1) and one value for the entire second bottom echo ‘attdecsv2’ (e2) were calculated (fig. 3). to obtain the average backscattering strength of the entire first second bottom echoes, respectively, attsvx and decsvx of the first and second echoes are converted into intensities. then, they are weighted using the attack and decay samples (na and nd), summarised, divided by (na + nd), and converted back into a db value using equations 3 and 4 (balk et al., 2011). (eq. 3) (eq. 4) in sonar 5 the number of attack samples (na) is 8, but the first sample that should be integrated into the decsv value is used for the calculation of neither attsv nor decsv. thus, the number of decay samples (nd) is 23 instead of 24, which must be considered in the formula when calculating e2 (poulain et al., 2011). sediment data acquisition (ground truthing) eleven sediment cores were taken from water depths between 2 and 15.7 metres. an elongated version (80 cm) of the ‘mondsee’ gravity corer by uwitech (niederreiter, 2012) was used to allow for a higher core penetration depth, as it was expected that the lower frequency is able to deeper penetrate soft sediment (dunbar et al., 1999). in addition to core sampling, 21 grab samples were taken using a petersen grab sampler (us environmental protection agency (epa), 2001). the material was put into a whirl paks® (nasco) sampling bag of two litres in volume until analysis. steep slopes and rocky sea beds proved to be harder to sample properly, as both sampling devices may tilt during sampling or rocks prevent the jaws from closing. unlike the uwitech corer, the grabber samples the top 15 cm of the deposits only and the fine fraction can be washed out partly during grabber recovery. sediment samples from both coring and grabbing were analysed for granulometry, loss on ignition (loi), organic carbon content, and wet bulk density (cores only) (harris et al., 2008). sediment sample fractions smaller than 63 µm were separated and additionally analysed for the contents of iron, phosphorus, and manganese. determination of granulometry was accomplished by wet sieving of the samples (deutsches institut fuer normung e.v., 2005), as the use of water guarantees that agglomerates dissolve and the particles are classified as the correct grain size. to obtain five classes of grain size, sieves with mesh widths of 2 mm, 500 µm, 250 µm, and 63 µm were used. the loi was measured according to din en 15169:2007, deutsches institut fuer normung e.v., (2007). between 2-5 g of each air-dried sample were filled into ceramic pots and weighed on a high-precision scale. they were then dried for 12 h at 105°c, cooled to room temperature in a desiccator, and weighed again. afterwards, the samples were burnt for two hours in a muffle furnace at 550°c, cooled down, and weighed again to determine the loi. as the sediment structure of the inner core is assumed not to be disturbed during core sampling, the wet bulk density of core samples in contrast to grab samples can be determined. this assumption was made, based on the visual findings that only the outer millimetres were affected by the impact of the tube and that the inner part of the 9 cm diameter was intact. the water saturated upper layer of the sediment settled again in most cases after the impact of the corer. a cylinder of fixed volume (17.6 cm³) was used to cut material out of the undisturbed inner core sample. from the weight of the fresh core material and the volume of the cylinder, the wet bulk density of the sediment was calculated. density was determined for each visual distinctive core layer and a vertical average was calculated for later correlation. due to the high share of fine particles in the sediment, no nco mm er cia l u se on ly hydro-acoustic sediment characterisation 99 chemical analyses concentrated on the silt and clay fractions. total carbon was determined using an eltra cs 2000 carbon sulfur determinator. the organic carbon content was calculated by subtracting the inorganic carbon content from the total carbon. total phosphorus was determined in the form of phosphate by molybdate blue analysis according to din 38405-11. the iron content was measured using flame atomic absorption spectrometry (perkin-elmer 1100b) and manganese content was measured using graphite furnace atomic absorption spectrometry (perkin-elmer simaa 6000). results sediment data (ground truthing) in all samples taken the silt and clay fraction was predominant. it accounted for an average weight of 74%, with a minimum of 15.5% and a maximum of 99.7%. however, about half of the sediment samples consisted almost completely (>90%) of silt and clay, whereas the other half contained considerable percentages of coarser fractions (fine, medium, and coarse sand and gravel). the loi ranged from 2.9% to 18.7%. wet bulk density values of the core samples varied from 1.08 g cm–³ to 1.61 g cm–³. the density of the cores correlated with various other sediment parameters, e.g. the share of particles smaller than 63 µm (r=-0.79), loss on ignition (r=-0.88), phosphorus (r=-0.80) and mn contents (r=-0.72). principal results of sediment analysis of all samples are given in tab. 3. hydro-acoustic survey to obtain a gapless 3d surface of the lake ground, the data of the bathymetric survey in the defined grid (100x50 m) were interpolated using the inverse distance weighting (idw) method (arcgis 10.2, esri). the bathymetric map shows a general depth gradient from the two major inflows to the dam with a depth maximum in front of the dam (fig. 4). the old river beds are still visible and represent the deepest depressions for each cross-section of the side arms. small branches have relatively shallow depths. the outer shape as well as the depth distribution illustrate the high morphometric complexity of the vossoroca reservoir. relationships between sediment and hydro-acoustic parameters to assess the use of the hydro-acoustic parameters for sediment characterisation and prediction, all combinations of relevant parameters were correlated. prior to the regression analyses, the sediment data sets were checked for normal distribution using the shapiro-wilk and kolmogorov-smirnov tests. with a confidence level of 5%, p-values of 0.77 for the kolmogorov-smirnov test and around 6.2 for the shapiro-wilk test were reached, which confirmed the normal distribution of the data sets. consequently, the analyses applied are statistically legitimate. tabs. 4 and 5 list the pearson r-values for the core samples and the grab samples. the 38 and 200 khz frequencies are shown next to each other for direct comparison. rvalues above the significance level are highlighted (core samples: p<0.05 for r=0.63, n=10; grab samples: p<0.01, r=0.55, n=21). for both sample types, the average performance of the lower frequency is clearly higher. in addition, the correlation with core samples reaches higher levels of conformance. the best detectable parameter is ‘density’. independently of the hydro-acoustic parameters, it reaches an average r-value of 0.62 and maximum r-values of 0.94. additionally, high correlations are reached between the particle composition and most hydroacoustic parameters. here, the best couple, % <63 µm with e1, reaches an r-value of 0.95. tab. 3. sediment key parameters from cores and grabs taken in vossoroca reservoir (n=32). mean max. min. sd cv depth (m) 8.8 15.6 2.1 3.6 0.4 loi (%) 12.1 18.7 2.9 4.4 0.4 org. carbon (%) 2.8 6.5 0.4 1.3 0.5 wet bulk density (g cm–3) 1.2 1.61 1.08 0.2 0.2 phosphorus (mg kg–1) 822 1,344 224 329 0.4 mn (mg kg–1) 341 488 190 71.3 0.2 fe (g kg–1) 42 68.3 11.4 12.6 0.3 proportion of particles <63 µm (%) 74 99.7 15.5 32.6 0.5 proportion of particles <250 µm >63 µm (%) 10.2 95.2 0.1 17.6 1.7 proportion of particles <500 µm >250 µm (%) 8.6 36.8 0 11.2 1.3 proportion of particles <2 mm >500 µm (%) 8.1 42.5 0 12.9 1.6 proportion of particles >2 mm (%) 1.7 15 0 3.2 1.9 cv, coefficient of variation. no nco mm er cia l u se on ly 100 s. hilgert et al. fig. 4. bathymetric map of the vossoroca reservoir. tab. 4. pearson r-values for the sediment and acoustic parameters based on the results obtained from the core samples of the vossoroca reservoir; in each case, the correlation for the best performing pulse length is shown only; statistical significance is given for r >0.63 p<0.05, n=10). density total p loi % <63 µm total c fe mn khz 200 38 200 38 200 38 200 38 200 38 200 38 200 38 e1’ -0.31 -0.92 -0.24 0.79 0.33 0.84 0.42 0.90 0.33 0.70 -0.42 0.69 0.19 0.72 e1 -0.64 -0.73 0.66 0.80 0.64 0.67 0.91 0.95 0.32 0.30 0.74 0.86 0.29 0.38 attdecsv1 -0.50 -0.93 0.26 0.81 0.50 0.84 0.55 0.94 0.38 0.69 -0.32 0.73 0.15 0.73 attsv2 0.65 -0.49 -0.76 0.27 -0.56 0.49 -0.82 -0.41 -0.25 0.74 -0.25 -0.56 -0.39 0.73 decsv2 -0.34 -0.94 -0.42 0.80 0.36 0.85 0.50 0.89 0.34 0.77 -0.45 0.66 -0.20 0.85 e2 0.39 -0.94 -0.54 0.77 0.31 0.86 -0.57 0.84 0.29 0.77 -0.60 0.61 -0.29 0.84 e1’/e1 -0.49 0.68 0.48 -0.48 0.39 -0.65 0.80 -0.50 -0.24 -0.72 0.78 0.53 -0.18 -0.77 e1/e2 0.67 -0.46 -0.72 0.38 -0.66 0.58 -0.91 -0.38 0.33 0.81 -0.81 -0.37 -0.33 0.80 no nco mm er cia l u se on ly hydro-acoustic sediment characterisation 101 even though some of the best correlations for grab samples reach the same range (best value: r=0.89), average values for certain parameters are clearly lower than the core sample results (tab. 5). e1/e2 is the best performing hydro-acoustic parameter at 200 khz frequency for the core samples with an average correlation of 63% over all pulse lengths and sediment parameters. for the 38 khz frequency, the best overall parameters are e1’, attdecsv1, and decsv2 with an average performance of 79%. for the correlation with the sediment characteristics obtained from grab samples, the e1 parameter produced the best results at both frequencies, with 68% at 200 khz and 67% at 38 khz. the best performing parameter couples for core samples averaged over all pulse lengths are e1/e2 with % <63 µm, with 74% at 200 khz frequency, and e1 combined with % <63 µm resulting in a correlation of 95%. for the grab samples, the best single sediment parameter is % <63 µm with 73% over all pulse lengths at 38 khz combined with the e1 parameter. the best detectable sediment parameter at 200 khz is the organic carbon content with 85% based on the attdecsv1 values. effect of pulse duration at all sampling positions, the listed pulse lengths (tab. 2) were used one after the other to investigate differences of the correlations with sediment parameters. since the pulse length has an influence on the hydroacoustic resolution, a variation in correlation performance can be expected (guillard et al., 2009). fig. 5 compares the effect of different pulse lengths on the r²values for selected hydro-acoustic features obtained using the 200 and 38 khz frequencies with core samples for ground truthing. for the 200 khz frequency, the shortest pulse length produces the best correlations independent of the sediment parameter. this accounts for the e1 as well as for the e1/e2 parameter, which includes energy patterns of the second bottom echo. the same result is shown by the e1 parameter for the 38 khz frequency, with the correlation being significantly better with shorter pulse length. different results are obtained for the e1/e2 parameter at 38 khz. all hydro-acoustic features measured at 38 khz, inclusive of the second bottom signal (attsv2, decsv2, e2, e1/e2), show better correlations for the longer pulse duration. cores versus grabs fig. 6 illustrates the overall result of correlations of hydro-acoustic values with sediment data obtained from core samples being significantly higher. this holds for both frequencies used in this survey. while the difference between core-related results and grab-related results is high for the attsv2-parameter, the difference is only marginal for the e1’ parameter. as the part of the signal processed for the e1’ parameter represents the first layer of the sediment only, the small difference between core and grab results is in accordance with the sampling depth of the grab sampling device (0-20 cm). both sampling techniques will produce comparable sediment samples for the upper sediment layer. in contrast to this, the results for all acoustic parameters, including deeper layers of the sediment, differ considerably between core and grab samples. in these cases, the correlation with the sediment parameters obtained from the core samples and, hence, from deeper layers fit well to the values of the hydro-acoustic survey. discussion combining two frequencies, various pulse lengths and multiple physical and chemical sediment parameters surveyed in the vossoroca reservoir in brazil resulted in a set of good and moderately correlated parameters. significant differences between the 200 and 38 khz frequencies and the tested sediment parameters were found. among these parameters were wet bulk density, granulometry, phostab. 5. pearson r-values for the sediment and acoustic parameters based on the results obtained from the grab samples of the vossoroca reservoir; in each case, the correlation for the best performing pulse length is shown only; statistical significance is given for r >0.55 (p 0.01, n=21). total p loi % <63 µm total c fe mn khz 200 38 200 38 200 38 200 38 200 38 200 38 e1’ 0.61 0.52 0.45 0.52 -0.47 0.72 0.89 0.29 0.51 0.43 -0.25 0.26 e1 0.75 0.75 0.72 0.73 0.72 0.79 0.80 0.66 0.73 0.73 0.36 0.39 attdecsv1 0.61 0.65 0.46 0.58 0.40 0.78 0.89 0.40 0.51 0.57 0.23 0.45 attsv2 0.38 -0.21 0.22 -0.21 -0.21 0.14 0.79 -0.52 0.32 -0.12 -0.32 0.17 decsv2 0.55 0.60 0.40 0.56 -0.37 0.73 0.89 0.39 0.47 0.53 0.21 0.36 e2 0.52 0.58 0.36 0.53 -0.35 0.72 0.88 0.37 0.44 0.51 0.18 0.36 e1’´/e1 0.73 0.47 0.64 0.40 0.63 0.34 0.84 0.61 0.72 0.52 0.29 0.26 e1/e2 -0.78 -0.37 -0.74 -0.43 -0.79 -0.28 -0.73 -0.57 -0.74 -0.40 -0.40 0.09 no nco mm er cia l u se on ly 102 s. hilgert et al. phorus content, organic c as well as iron and manganese contents. comparison of the correlations of sediment parameters obtained from core or grab sampling revealed that core samples are clearly better suited for ground truthing. additionally, analysis of the correlation performance of a range of tested pulse lengths showed that most echo features produce better results at shorter pulse lengths. in additional experiments, depth dependence on seabed classification was investigated, but since the results did not show any clear tendencies, they are not presented in this article. as most of the available literature does not focus on the acoustic determination of sediment properties like carbon content or total p content, data available for comparison are scarce. still, the results can be compared partly with the results from anderson and pacheco (2011). comparable are the r-values at 200 and 38 khz for the acoustic parameters e1’/e1 and e1/e2 and the sediment parameters clay content, loi, and total p. the results show similar tendencies. while the 200 khz frequency can detect the granulometric properties, the 38 khz fails to reach a significant level of correlation (tab. 4). the loi and total p content are largely described by the e1/e2 parameter, but not by the e1´/e1. this is in conflict with the results fig. 5. comparison of r2-values obtained with different pulse lengths for selected hydro-acoustic features using 200 and 38 khz frequencies and core samples for ground truthing. no nco mm er cia l u se on ly hydro-acoustic sediment characterisation 103 of anderson and pacheco (2011). thirty eight khz r-values for loi and p content show different patterns, as the e1´/e1 parameter produces better results compared to the e1/e2. in general, the p content could be correlated with high r-values, which contradicts the results presented by anderson and pacheco, 2011. even if the exact values differ from the results of anderson and pacheco (2011), it can be stated that the granulometric properties of the sediment can be related directly to the acoustic parameters. taking into account other acoustic parameters (e1’, e1, e2) and their correlations with granulometric features, even higher correlations of up to 93 % were found (tab. 4). the best correlations were obtained for the wet bulk density and the share of particles <63 µm. the negative correlation with the wet bulk density can be attributed to the presence of gas bubbles in the sediment matrix. bubbles cause a lower wet bulk density and a higher reflection intensity, since accumulations of gas bubbles represent strong reflectors of sound impulses (anderson et al., 1998; anderson and martinez, 2015). this assumption fits to the strongly positive correlation with the share of the silt and clay fraction. finer sediment layers are prone to have an increased content of organic carbon and, hence, a higher potential productivity of gas in the sediment. averaging of the correlation results for silt and clay at 38 khz in anderson and pacheco, 2011 yielded -0.31, which is very close to the r-value of -0.28 obtained here. additionally, the correlations for total p (r=-0.37, 38 khz) and loi (r=-0.43, 38 khz) reproduce the published results well. it can be concluded that i) correlation and ii) the coefficient values yielded similar results for comparable parameter combinations. in contrast to literature, this study shows highly significant correlations for total p (r >0.93) (attdecsv1, decsv2, e2) and various sediment attributes (tab. 4). in general, the results at 38 khz frequency produce slightly higher correlation coefficients than at 200 khz frequency. for the grain size parameters, the coefficients have similar values (van walree et al., 2005). this agreement between the hydro-acoustic response of the phosphorus content and the particle size distribution seems to be reasonable, since most phosphorus species are primarily bound to the silt and clay fraction. furtherfig. 6. comparison of the correlation results of the eight hydro-acoustic parameters with % <63 µm for both frequencies and cores and grabs. the bars show the average r2-value while the error bars represent the maximum and minimum values from the different configurations. no nco mm er cia l u se on ly 104 s. hilgert et al. more, the density of the sediment was clearly better detected by the 38 khz frequency, reaching correlations of 95% (bentrem et al., 2002). the better correlation is due to the position of the density measurements in the cores. sampling material was extracted from around 10 cm below the sediment / water interface, which reduces the potential share of the sediment volume represented in the returned 200 khz acoustic signal, because the 200 khz frequency is not able to penetrate the sediment very deeply. especially a shorter pulse length reduces the signal-relevant sediment depth. grab sampling for ground truthing is a widely spread standard method. however, the presented results show a significant difference between ground truthing results from cores and grabs (fig. 6). this especially applies to physical sediment parameters, but also to the loi and organic c. differences between cores and grabs are highest for all correlations, including second bottom-echo features, and lowest for the e1’ parameter. the results can be explained by the fact that the sediment relevant to the e1’ parameter is the same for both sampling approaches, while the sediment volume ensonified during the e1 phase (4 * τ=up to 3 m) cannot be sampled with a grab sampler. since the influence of the sediment volume properties increases from the first to the second bottom echo, the difference between cores and grabs is enhanced. here, the internal sediment layering plays an important role (ostrovsky and tęgowski, 2010). due to the higher penetration depth of the 38 khz frequency, the influence of the sediment volume is even stronger. for this reason, divergence between core and grab results is larger than at the 200 khz frequency. additionally, it must be stated that the sediment type may play a major role regarding the influence of the sampling method. in the presented case, sediment could be properly sampled with both techniques. however, higher shares of coarse material (e.g., gravel) or lower cohesion of the sample volume may lead to biased results between core and grab samples. regarding the discrepancy between the ensonified footprint area and the surface sampled by the grabber or sediment corer, a potential bias might occur. during this study, the mean sampling depth was 8.8 m. for the 200 khz frequency, it caused a footprint of ~3.8 m² and for the 38 khz frequency of ~12.2 m². it can be assumed that, except for very narrow side arms, the sediment composition does not change significantly within 2-3.5 m. consequently, the relatively small sampling area of the corer and grabber can be neglected until a certain depth. for sediment investigations at greater depth, a higher number of sediment samples per sampling location seems to be advisable. based on the results presented, the echo sounder settings can be optimised for sediment characterisation. specific settings can be chosen for the sediment parameter to be determined. in dependence to the used frequency, a shorter pulse length will produce better classification results. optimal configurations allow for a good estimation (r=0.9-0.95) of the granulometry, bulk density or even loi. lower but still significant results were achieved for total p (r=0.81). additionally, the results show that core sampling produces significantly better results. the presented results prove that under the given conditions, sediment characteristics can be detected. especially in morphometrically complex systems, this will ensure improved sediment monitoring and, later on, better management strategies, as they are based on sediment information. here, information on the sediment type (granulometry) and potential organic content or even phosphorus content can help in planning sediment disposal. sediment information can also be used in the context of methane production, as the productivity of the sediment is linked to the amount of organic carbon in the sediment (sobek et al., 2012). conclusions • the exact configuration of the echo sounder has a strong influence on potential sediment characterisation and sediment feature correlation. • shorter pulse lengths produce better ground truthing results. • core samples produce significantly better results for ground truthing than grab samples, especially at the lower frequency and for decay-related hydro-acoustic features. • in addition to physical parameters, chemical parameters can be included in hydro-acoustic sediment characterisation. • the produced information may significantly improve sediment management even in large-scale reservoirs. acknowledgments many thanks go to the ufpr, dhs research team with prof. fernandes, prof. bleninger, and julio werner as well as to the technical staff of the dhs, who supported the research activities. we also thank prof. helge balk from the university of oslo for support with the sonar5-pro software. for the gps support, we thank prof. claudia krueger from the departamento de geomática, ufpr. additional thanks go to prof. fernandes, prof. bleninger, and prof. helge balk for reviewing this article. references amiri-simkooei ar, snellen m, simons dg, 2011. principal component analysis of single-beam echo-sounder signal features for seafloor classification. ieee j. ocean. eng. 36:259-272. no nco mm er cia l u se on ly hydro-acoustic sediment characterisation 105 anderson al, abegg f, hawkins ja, duncan me, lyons ap, 1998. bubble populations and acoustic interaction with the gassy floor of eckernförde bay. cont. shelf res. 18:18071838. anderson jt, van holliday d, kloser r, reid dg, simard y, 2008. acoustic seabed classification: current practice and future directions. ices j. mar. sci. 65:1004-1011. anderson ma, martinez d, 2015. methane gas in lake bottom sediments quantified using acoustic backscatter strength. j. soil. sediment. 15:1246-1255. anderson ma, pacheco p, 2011. characterization of bottom sediments in lakes using hydroacoustic methods and comparison with laboratory measurements. water res. 45:4399-4408. balk h, lindem t, sánchez-carnero n, 2011. sonar4 and sonar5 post processing systems. operator manual version 6.0.1. extension for seabed classification tool. bentrem fw, sample j, kalcic mt, duncan me, 2002. highfrequency acoustic sediment classification in shallow water. oceans-ieee 1:7-11. burczynski j, 1999. bottom classification. available from: www.biosonicsinc.com deutsches institut fuer normung e.v., 2005. aggregates test methods. determination of particle size distribution by wet sieving. beuth verlag gmbh. deutsches institut fuer normung e.v., 2007. characterization of waste. determination of loss on ignition in waste, sludge and sediments. beuth verlag gmbh. dunbar ja, allen pm, higley pd, 1999. multifrequency acoustic profiling for water reservoir sedimentation studies. j. sediment. res. 69:521-527. freitas r, sampaio l, oliveira j, rodrigues am, quintino v, 2006. validation of soft bottom benthic habitats identified by single-beam acoustics. mar. pollut. bull. 53:72-79. friedl g, wüest a, 2002. disrupting biogeochemical cycles consequences of damming. aquat. sci. 64:55-65. guillard j, godlewska m, colon m, doroszczyk l, dlugoszewski b, 2009. standartization of hydroacoustic methods effect of pulse duration. in: proc. 3rd int. conf. and exhibition of underwater acoustic measurements: technologies & results, nafplion, greece. hamilton lj, parnum i, 2011. acoustic seabed segmentation from direct statistical clustering of entire multibeam sonar backscatter curves. cont. shelf res. 31:138-148. harris mm, avera we, abelev a, bentrem fw, bibee ld, 2008. sensing shallow seafloor and sediment properties. recent history. oceans-ieee 2008(suppl.):1-11. niederreiter r, 2012. uwitech sampling equipments. available from: www.uwitec.at/html/frame.html odhiambo bk, boss sk, 2004. integrated echo sounder, gps, and gis for reservoir sedimentation studies: examples from two arkansas lakes. j. am. water resour. as. 40:981-997. orlowski a, 1984. application of multiple echoes energy measurements for evaluation of sea bottom type. oceanologia 1984:61-78. ostrovsky i, tęgowski j, 2010. hydroacoustic analysis of spatial and temporal variability of bottom sediment characteristics in lake kinneret in relation to water level fluctuation. geomar. lett. 30:261-269. parnum i, siwabessy j, gavrilov a, parsons m, 2009. a comparison of single beam and multibeam sonar systems in seafloor habitat mapping, p. 155-162. in: proc. 3rd int. conf. and exhibition of underwater acoustic measurements: technologies & results, nafplion, greece. poulain t, argillier c, gevrey m, guillard j, 2011. identifying lakebed nature: is it feasible with a combination of echosounder and sonar5-pro? adv. oceanogr. limnol. 2:49-53. preston jm, 2009. automated acoustic seabed classification of multibeam images of stanton banks. app. acoust. 70:12771287. republic of brazil canambra engineering consultant, united nations development programme,1969. power study of south brazil. sobek s, delsontro t, wongfun n, wehrli b, 2012. extreme organic carbon burial fuels intense methane bubbling in a temperate reservoir. geophys. res. lett. 39:l01401. tęgowski j, 2005. acoustical classification of the bottom sediments in the southern baltic sea. quatern. int.130:153-161. us environmental protection agency, 2001. measurement and monitoring technologies for the 21st century. available from: https://clu-in.org/programs/21m2/ van walree pa, tęgowski j, laban c, simons dg, 2005. acoustic seafloor discrimination with echo shape parameters: a comparison with the ground truth. cont. shelf res. 25:2273-2293. wcd (world commission on dams), 2000. dams and development: a new framework for decision-making. earthscan publications, london. no nco mm er cia l u se on ly layout 1 introduction cladocera are one of the most important components of zooplankton in lacustrine ecosystems. this group is also well-known to promptly respond to environmental changes (jeppesen et al., 2001). previous studies have highlighted that temperature and altitude are among the main factors determining cladocera communities (campanelli, mortari and henry, 2016; green, 1995; padhye et al., 2016; sweetman, 2010; zawiska et al., 2015). in addition, these studies have shown comparable sensitivities to changes in lake water trophic state, water depth, conductivity and ph (korhola and rautio, 2001). these responses to environmental variables makes cladocera one of the most powerful paleolimnological indicators and for this reason they have been widely used for the reconstruction of past lacustrine environments (bjerring, 2007; kienast et al., 2011; korosi, 2012; luoto et al., 2009; mirosław-grabowska and zawisza, 2014; nováková et al., 2013 sarmaja-korjonen and hyvärinen, 2008; schmidt, 2000; szeroczyńska, 1991; szeroczyńska et al., 2007; szeoczyńska and zawisza, 2011 a,b). subfossil remains of cladocera are commonly used in paleolimnological studies in eurasia and america (bjerring, 2007; mirosław-grabowska and zawisza, 2014; paterson, 1994; szeroczyńska et al., 2007). however, only a few paleolimnological studies based on water fleas were conducted in central america (cuna et al., 2014; szeroczyńska et al., 2015; zawisza et al., 2012, 2014, 2016). so far, the majority of studies focused on the living cladocera and their taxonomy (elías-gutiérrez et al., 1999, 2006, 2008; sinev, 2015; sinev and silva-briano, 2012; sinev and zawisza, 2013). the knowledge of cladocera species composition, distribution and ecology in freshwater lakes of central america is crucial for the development of reliable bioindicators, and for paleolimnological investigations. the main objective of the present study was to analyze and identify subfossil cladocera species in northern central america, in particular in guatemala, el salvador and honduras, and to explore their relationship to lake environmental characteristics such as altitude, lake area, depth, ph, transparency, conductivity and dissolved oxygen. the obtained results will provide the basis for the reconstruction of regional and global climatic and envoronemntal changes at long-term scale. advances in oceanography and limnology, 2016; 7(2): 151-162 article doi: 10.4081/aiol.2016.6266 this work is licensed under a creative commons attribution-noncommercial 4.0 international license (cc by-nc 4.0). ecology of cladocera species from central america based on subfossil assemblages marta wojewódka,1* edyta zawisza,1 sergio cohuo,2 laura macario-gonzález,2 antje schwalb,2 izabela zawiska,3 liseth pérez4 1institute of geological sciences, research centre in warsaw, polish academy of sciences, twarda 51/55, 00818 warsaw, poland; 2institut für geosysteme und bioindikation, technische universität braunschweig, langer kamp 19c, 38106 braunschweig, germany; 3institute of geography and spatial organization, polish academy of sciences, twarda 51/55, 00818 warsaw, poland; 4instituto de geología, universidad nacional autónoma de méxico (unam), ciudad universitaria, 04510, ciudad de méxico, méxico *corresponding author: m.wojed@twarda.pan.pl abstract cladocera species composition was analyzed in surface sediments of 29 lakes in central america (guatemala, el salvador and honduras). the material studied was collected with an ekman grab in autumn 2013 from lakes located in lowland, highland and mountain regions. the study revealed high variability in qualitative and quantitative composition of subfossil cladocera. a total of 31 cladocera species (5 planktonic and 26 littoral) were identified, as well as 4 morphotypes that could not be identified (nrr 1-4). planktonic bosminidae and daphniidae were the most abundant families. daphniidae were restricted to water bodies in mountain regions, whereas bosminidae were widely distributed in lakes with different abiotic conditions. moreover, bosminidae species also occurred in highly mineralized waters (> 900 µs cm–1). the great majority of the identified cladocera species belonged to the littoral family chydoridae. chydorus cf. sphaericus was the most common species (found in 20 lakes), which probably reflects its tolerance to a wide spectrum of habitat conditions. cluster analysis discriminated 6 groups of cladocera species with a high correlation level within groups (≥0.8), which showed different types of correlation with lake characteristics and environmental variables. canonical correspondence analysis (cca) showed that altitude and secondly water electrical conductivity were the most important drivers of cladocera species composition in the region studied. furthermore, cca analysis indicated lowland lakes with low water transparency were also characterized by peculiar species assemblages. key words: subfossil cladocera; central america; waterfleas ecology; guatemala; el salvador; honduras. received: september 2016. accepted: november 2016. no n c om me rci al us e o nly 152 m. wojewódka et al. methods study sites the study included 29 waterbodies located between 13° and 17° n geographical latitude and from 87° to 91° w longitude in guatemala, el salvador and honduras (fig. 1). the study area is located in the central american bioregion of the american neotropical ecozone, which is characterized by two distinct seasons: a dry winter and a moist summer season (taylor and alfaro, 2005). the subtropical ridge (belt of high atmospheric pressure) from the northern atlantic is controlling the climate of the region, whereas the intertropical convergence zone (itcz) and polar fronts of mid-latitude origin have a secondary impact (taylor and alfaro, 2005). one of the main factors affecting the microclimate of the area is topography (karmalkar et al., 2011; taylor and alfaro, 2005). the considerable differences in mean annual temperature and precipitation which characterize the entire region mainly depend on altitude. the guatemalan lowlands are characterized by a mean annual temperature of 25-28°c, whereas temperatures in highland and mountain regions are around 12-15°c (atlas climatologico for 1928-2003, the national institute for seismology, volcanology, meteorology and hydrology of guatemala; www.insivumeh.gob.gt). annual precipitation in the area varies considerably during the year and ranges from <1000 mm in the plains to >2500 mm in the mountains (taylor and alfaro, 2005). the lakes studied are located at altitudes ranging from sea level (el muchacho at 3 m asl) to high altitude lakes (magdalena and chicabal at 2863 and 2726 m asl, respectively, tab. 1). lake area and depth also were quite variable. the largest lake were atitlán and yojoa (125 and ~79 km2, respectively, tab. 1), while several lakes have an area <0.10 km2, such as madre vieja (0.10 km2), verde fig. 1. map of the study area showing the location of lakes sampled in central america, i.e. guatemala, honduras and el salvador. numbers indicate the names of the lakes: 1, atitlán; 2, calderas; 3, chicabal; 4, comandador; 5, el muchacho; 6, el pino; 7, el rosario; 8, grande; 9, ipala; 10, lachuá; 11, las pozas; 12, magdalena; 13, sacnab; 14, salpetén; 15, quexil; 16, chiligatoro; 17, jucutuma; 18, madre vieja; 19, ticamaya; 20, yojoa; 21, apastepeque; 22, aramuaca; 23, chanmico; 24, el espino; 25, jocotal; 26, los negritos; 27, metapan; 28, olomega; 29, verde. no n c om me rci al us e o nly subfossil cladocera of central america 153 (0.10 km2), chiligatoro (0.04 km2), el rosario (0.02 km2), magdalena (0.01 km2) (tab. 1). the study covered both very shallow (≤3 m depth, such as comandador, el muchacho, el rosario, grande, magdalena, jucutuma, ticamaya, jocotal, los negritos, olomega), and very deep lakes (>300 m), such as lachuá (a 378 m deep karstic lake), atitlán (a 340 m deep crater lake) and chicabal (330 m, tab. 1). sampling and analyses sediments were collected in autumn 2013 using an ekman sediment sampler. the geographical location of the lakes was determined by a handheld navigator (gpsmap 60c). water transparency was determined using a secchi disk, and the maximum lake depth was identified using a portable depth sounder (echosounder eagle mach 1). physical and chemical parameters of the surface waters (i.e., ph, conductivity, dissolved oxygen) were measured with a wtw multi set 350i multiparametric probe. the area of each lake (tab. 1) was calculated by the measure tool of google earth. lake sediment samples were collected both from the littoral as well as from the open-water (pelagic) zone. in eight shallow lakes, where there was no pelagic zone (i.e., comandador, el muchacho, el rosario, grande, jucutuma, ticamaya, jocotal, los negritos), samples were collected only from the littoral zone (tab. 1). lakes el pino, ipala, quexil, madre vieja, chanmico, el espino, metapan and verde, were sampled only in the profundal zone due to the shape of the lake basin. crater lake aramuaca (107 m depth) was sampled only in the littoral zone (tab. 1). tab. 1. location, morphometry and selected limnological variables measured at water surface at the waterbodies sampled in central america. id name country geographic altitude area samples zmax secchi ph conductivity o2 hcl coordinates n w (m asl) (km2) l p (m) (m) (µs cm–1) (mgl–1) 1 atitlán guatemala 14.6837 91.2239 1556 125.00 + + 340 4.2 8.9 442 5.7 + 2 calderas 14.4117 90.5913 1790 0.35 + + 26 2.9 9.2 100 4.9 3 chicabal 14.7875 91.6561 2726 0.21 + + 330 1.6 9.0 12 5.4 4 comandador 13.9600 90.2544 20 0.65 + 1.7 0.5 7.4 251 5.6 5 el muchacho 13.8892 90.1918 3 0.36 + 2 0.4 9.1 439 5.9 6 el pino 14.3447 90.3941 1038 0.64 + 18 2.8 8.3 100 2.0 7 el rosario 16.5255 90.1601 126 0.02 + 3 0.2 7.2 987 1.7 + 8 grande 13.8903 90.1703 5 0.95 + 2 0.2 7.4 110 3.7 9 ipala 14.5571 89.6394 1495 0.56 + 25 3.6 8.0 100 2.8 10 lachuá 15.9184 90.6732 170 4.00 + 378 4.4 7.9 906 4.8 + 11 las pozas 16.3432 90.1660 152 2.16 + + 35 3.1 8.3 277 5.4 + 12 magdalena 15.5426 91.3956 2863 0.01 + + 3 2.8 8.8 331 6.2 + 13 sacnab 17.0583 89.3725 170 4.28 + + 9 1.8 9.0 412 6.1 + 14 salpetén 16.9815 89.6755 105 2.77 + + 32 1.7 7.0 4520 5.8 + 15 quexil 16.9231 89.8099 120 2.20 + 32 2.7 8.5 204 4.5 16 chiligatoro honduras 14.3756 88.1830 1925 0.04 + + 5.5 0.9 7.4 100 1.9 17 jucutuma 15.5123 87.9028 27 4.34 + 2 2.0 7.3 100 1.0 18 madre vieja 14.3569 88.1376 1866 0.10 + 3.4 0.9 8.5 100 2.7 19 ticamaya 15.5506 87.8897 17 2.91 + 2 0.8 7.2 100 1.6 20 yojoa 14.8606 87.9847 639 79.70 + + 22 3.3 8.3 100 2.7 21 apastepeque salvador 13.6925 88.7448 509 0.38 + + 47 6.1 8.6 100 2.8 22 aramuaca 13.4294 88.1065 76 0.40 + 107 6.7 8.4 100 4.1 23 chanmico 13.7786 89.3541 477 0.78 + 51 0.9 9.2 100 3.4 + 24 el espino 13.9530 89.8652 689 0.99 + 5.5 0.4 8.5 85 6.6 25 jocotal 13.3371 88.2519 26 8.70 + 3 1.5 8.0 595 3.0 26 los negritos 13.2831 87.9370 102 0.69 + 2 0.3 9.2 40 5.2 27 metapan 14.3094 89.4655 450 16.00 + 6 0.2 8.4 255 3.1 28 olomega 13.3072 88.0551 66 25.20 + + 2.5 0.9 7.7 105 2.5 29 verde 13.8915 89.7872 1609 0.10 + 12 2.7 7.5 83 4.5 l, littoral zone; p, pelagic zone; zmax, maximum lake depth; hcl, reaction with acid during preparation. the world geodetic system of 1984 (wgs84) datum was used for the cladocera samples. no n c om me rci al us e o nly 154 m. wojewódka et al. sediment samples were placed in plastic bottles (100 ml) immediately after collection, and stored under refrigeration. samples were then transported to the paleolimnology laboratory of universidad national autonomous de mexico (unam), where subsamples for subfossil cladocera analysis were obtained and preserved in cooling condition. subfossil cladocera were analyzed at the bioindicator laboratory at the warsaw research center of the institute of geological sciences, the polish academy of sciences. each sample consisting of 1 cm3 of wet sediment was prepared according to standard methods (frey, 1986). sediments were treated with 10% hcl to eliminate carbonate and then heated in 10% koh for 30 min. the residue was centrifuged and sieved through a 38 µm sieve, transferred into a test tube and filled up to 5 cm3 with distilled water. samples were stained with safranin. three to four slides obtained from of 0.1 cm3 liquid suspension each were examined at the microscope for each sample. the identification of cladocera taxa followed elías-gutiérrez et al. (2006, 2008), hudec (2000), kotov (2009), kotov et al. (2003), rajapaksa and fernando (1987), rey and vasquez (1986), sinev (1998, 2001, 2015a, 2015b), sinev and dumont (2016), sinev and zawisza (2013), van damme et al. (2011). the cladocera relative abundance diagram was prepared using the c2 program (juggins 2005, 2007). in order to determine the relation between species distribution and lake characteristics (area, altitude, maximum depth) and water variables (ph, secchi transparency, dissolved oxygen), cluster analysis and canonical correspondence analysis (cca; ter braak, 1986) were performed. cluster analysis is based on the similarity measure performed on qualitative data (hammer et al., 2011), while cca is a combination of mutual averaging algorithm with a multiple regression (ter braak, 1986) which performed on combined quatitative cladocera and environmental data. statistical analyses were conducted using the past software (hammer et al., 2001). results environmental variables in general, the surveyed lakes were characterized by relatively low water transparency, with exception of lakes aramuaca and apastepeque, which showed secchi depth >6 m (tab. 1). secchi depth values ranging between 2.74.4 m were measured in 10 lakes, i.e. lachuá, atitlán, ipala, yojoa , las pozas , calderas, el pino, magdalena, quexil and verde (tab. 1). the waters of lake jucutuma (2.0 m depth) were characterized by visibility reaching the bottom, while four lakes were characterized by clarity of 1.5-1.8 m (sacnab, salpetén, chicabal, jocotal). the other lakes (comandador, el muchacho, el rosario, grande, chiligatoro, madre vieja, ticamaya, chanmico, el espino, los negritos, metapan, olomega) were characterized by secchi disk visibility <1.0 m (tab. 1). the lakes investigated did not show considerable differences in ph values, which ranged from neutral (7.0 in lake salpetén) to alkaline (9.2 in lake chanmico and lake los negritos, tab. 1). circumneutral water ph was recorded for lakes el rosario, ticamaya, jucutuma, comandador, chiligatoro, grande, olomega, verde and lachuá. waters of five lakes were strongly alkaline (ph ≥9.0), i.e., sacnab, el muchacho, calderas, los negritos, and chanmico. the other 14 lakes were characterized by slightly alkaline waters, with ph values ranging from 8.0 to 9.0 (tab. 1). electric conductivity ranged from 12 µs cm–1, measured at lake chicabal, to 4520 µs cm–1 in lake salpetén (tab. 1). high conductivity was recorded in the waters of el rosario and lachuá (tab. 1). values between ~400 and 600 µs cm–1 were determined in lakes jocotal, atitlan, el muchacho, and sacnab (tab. 1). in the remaining lakes this parameter was <350 µs cm–1 (tab. 1). ten of the sampled lakes were characterized by the dissolved oxygen content of over 5.0 mg l–¹, i.e., el espino, magdalena, sacnab, el muchacho, salpetén, atitlán, comandador, las pozas, chicabal and los negritos (tab. 1). the amount of dissolved oxygen in nineteen lakes ranged from 1.0 mg l–¹ (jucutuma) to 4.9 mg l–¹ (calderas). the lowest dissolved oxygen concentration (<2.0 mg l–¹) were recorded in lakes chiligatoro, el rosario, ticamaya and jucutuma (tab. 1). subfossil cladocera a total of 31 cladocera taxa belonging to three families – daphniidae, bosminidae and chydoridae – were found in the sediments of the lakes studied. some cladocera remains that could not be identified to species level belong to four different morphotypes and were referred to as “not recognized remains” (nrr) 1, 2, 3 and 4, respectively (fig. 2). species relative abundances are presented in fig. 3. frequency (n), mean and maximum relative abundance of each cladocera species are collated in tab. 2. the species richness per lake ranged from one in lake calderas to 12 species (lakes atitlán and verde), and included both planktonic and littoral forms. only in the sediments of the deep aramuaca crater lake no subfossil cladocerans remains were found. planktonic species were represented by five taxa, i.e., bosmina longirostris (o.f.müller, 1785), bosmina (e.) longispina (leydig, 1860), bosmina (e.) coregoni (baird, 1857), group of daphnia longispina (o.f. müller, 1785), group of daphnia pulex (leydig, 1860), which occurred in 17 lakes (fig. 3). cladocera communities from lakes el rosario and calderas were entirely dominated by planktonic species (fig. 3), while the share of bosminidae and/or no n c om me rci al us e o nly subfossil cladocera of central america 155 daphniidae exceeded 90% in lakes ipala, el pino and apastepeque (fig. 3). pelagic species were dominant (6088% of total cladocera abundance) in lakes salpetén, yojoa, el espino, atitlán, sacnab, chicabal, lachuá (fig. 3). the remaining lakes (16) were dominated by littoral species of chydoridae, which also represented the most species-rich group. in fact, 26 littoral cladocera species and four nrr (fig. 3 and tab. 2) were identified in the sediments studied, with the nrr most likely belonging to also to littoral chydoridae. chydorus cf. sphaericus (o.f. müller, 1776) (20 lakes), alona glabra (sars, 1901) (ovalona glabra, following sinev, 2015a) (13 lakes) and cladocera comparable to (13 lakes) and alona quadrangularis type (o.f. müller, 1776) (12 lakes) were the most common taxa (fig. 3 and tab. 2). in five lakes (las pozas, quexil, verde, chiligatoro, magdalena), both littoral and planktonic species were present. the share of littoral species ranged from 60% to 80% in lakes las pozas, quexil, verde and chiligatoro. in the sediments of lakes las pozas and quexil, eight and nine littoral species were identified, respectively, and two planktonic species were detected. in las pozas, alona ossiani (sinev, 1998), was the most abundant littoral species (approx. 20%), whereas in quexil – chydorus cf. sphaericus (ca. 15%) and alona quadrangularis type (ca. 15%, fig. 3) were dominant. lake verde and lake chiligatoro were characterized by the presence of 12 and 6 species, respectively. chydorus cf. sphaericus (ca. 25%) and group of anthalona verrucosa (sars, 1901) (ca. 15%) were the dominant species of the littoral zone in lake verde, while alona ossiani (ca. 35%) and simocephalus sp. (schoedler, 1858) (ca. 20%) were the most abundant species in lake chiligatoro (fig. 3 and tab. 2). in lake magdalena, the deep-water species accounted for a minor portion of cladocerans, and chydorus cf. sphaericus was the dominant species (ca. 90%). planktonic taxa were completely absent in the sediments fig. 2. unidentified subfossil remains (nrr) of cladocera. a) probably headshield of euryalona sp. b) headshield nrr1. c) postabdomen of leydigiopsis ornata. d) probably headshield of leberis sp. e) nrr3, probably shell of cladocera from group of coronatella monacantha. f) shell nrr4. scale bars: 500 µm. no n c om me rci al us e o nly 156 m. wojewódka et al. of eleven lakes (fig. 3). lakes comandador, los negritos, ticamaya and jocotal were dominated by chydorus cf. sphaericus, which represented from 36% to 50% of the cladocera communities. this taxon was also common in madre vieja (15%), metapan (23%) and olomega (16%). in lake madre vieja, alona ossiani was the most abundant species (ca. 42%), camptocercus dadayi (stingelin, 1913) was the dominant species in metapan (ca. 31%), and nrr1 (ca. 32%) (figs. 2 and 3) were dominant in olomega. in lake jucutuma, acroperus sp. (baird, 1843) (ca. 22%) and camptocercus dadayi (ca. 22%) were dominant, whereas alona quadrangularis type (ca. 43%) and alona glabra (ca. 29%) dominated in lake grande. lake el muchacho and lake chanmico were characterized by low species diversity (3) and low frequency of individuals. group of leydigia acanthocercoides (fischer, 1854) was dominant in lake el muchacho (ca. 52%), whereas leberis sp. (smirnov, 1989) was the most prominent species in lake chanmico (50%, fig. 3). the largest number of subfossil cladocera remains was determined for lakes sacnab (22,000 ind. cm–3) and ipala (21,000 ind. cm–3), where open-water species dominated. the maximum number of water fleas in the waterbodies dominated by littoral forms was ca. 10,000-11,000 ind. cm–3, as recorded in madre vieja, quexil and verde. statistical analysis the “cophenetic coefficient” in the cluster analysis was 0.8009 and well reflected the similarities within the data. the dendrogram shown in fig. 4 (see also tab. 3) defined six cladocera groups. the highest correlation coefficient between species was determined for alona rustica (scott, 1895) (group of flavalona rustica, following sinev and dumont, 2016) and simocephalus sp. (0.99, group 6), and for daphnia pulex-group and bosmina longirostris (0.94, group 1, fig. 4a and tab. 3). the species bosmina (e.) coregoni, bosmina (e.) longispina, alona quadrangularis fig. 3. relative abundances (%) diagram for the subfossil cladocera species identified in surface sediments from freshwater ecosystems of central america. lakes are ordered by decreasing contribution of planktonic forms. nrr1, unidentified remains type 1; nrr2, unidentified remains type 2, probably headshield of leberis sp.; nrr3, unidentified remains type 3, probably of coronatella of monacantha-group); nrr4, unidentified remains type 4. no n c om me rci al us e o nly subfossil cladocera of central america 157 type, alona glabra and leberis sp. (group 2) resulted to be correlated at the similarity level of 0.92. in the three other groups the correlation coefficient between species ranges between 0.8 and 0.86 (fig. 4a). the cca analysis was conducted using the species included the six distinguished clusters and an additional set of other cladocera species that were not included in any cluster. due to their rarity and low contribution to total cladocera remains, nrr were not included in the analysis. the cca ordination biplot is presented in fig. 4b. arrows show environmental variables, and their length and direction indicate the strength and direction of their influence on the species composition (ter braak, 1986). axis 1 and 2 explained almost 80% of the variability (axis 1=49.98%, axis 2=29.29%). altitude, transparency, ph and the maximum depth were positively correlated with axis 1, while conductivity and lake area were negatively correlated with the first canonical axis (fig. 4b). the location of group 1 in quadrant i, and of group 3, group 6 and the daphnia longispina-group in quadrant iv underlines the positive correlation of the abundances of these species groups with altitude and negative correlation with conductivity. moreover, the occurrence of graptoleberis testudinaria (fischer, 1848) and group 1 also seem to be strongly determined by the secchi disk visibility (fig. 4b). species and groups of cladocera located in quadrants ii and iii were correlated negatively with altitude (fig. 4b) and positively with conductivity and lake area (fig. 4b and tab. 3). group 2 and alonella excisa (fischer, 1854) were placed in the 2nd quarter. species of group 4 and 5 and epheromorphus sp. (frey, 1982), leydigia louisi louisi (jenkin, 1934), leydigiopsis ornata (daday, 1905), notalona sculpta (sars, 1901) were located in the 3rd quarter. according to fig. 4b, these species dominated in lowlands and in the waters with low visibility (negative correlation with secchi disk visibility and altitude). chydorus cf. sphaericus was located between quadrants iii and iv, which may indicate a weak relation with all of the considered environmental variables. discussion thirty-one cladocera species were identified in surface sediments of the lakes studied. the number of species found in individual lakes was generally small. in fact, the maximum number of species found in lakes atitlán and verde was only 12. compared to european lakes, which are normally inhabited by over 30 species (dumont, 1994), the species diversity of cladocera was low in the considered region of central america. these findings seems to confirm that the number of cladocera species found in one waterbody in the neotropics is in general low and usually ranges from several to over a dozen species (cuna et al., 2014; peréz et al., 2013; zawisza et al., 2012). the differences in the number of cladoceran species between the continents may result from species competition (feniova et al., 2011; shurin, 2000), fish or invertebrate predation (korhola and rautio, 2001), as well as from different geological history of the area (korhola and rautio, 2001). the identified cladocera included both littoral and pelagic species. planktonic species of daphniidae and bosminidae were the most abundant. remains of the daphnia pulex-group and the daphnia longispina-group were mainly associated to lakes located at altitudes above tab. 2. number of occurrences, maximum and mean percentage contribution of cladoceran species identified in the sediments of the surveyed waterbodies. group or species n max mean (%) (%) bosmina longirostris 9 100.0 43.8 bosmina (e.) coregoni 8 31.3 17.0 bosmina (e.) longispina 10 80.0 36.2 daphnia longispina-group 8 67.9 25.6 daphnia pulex-group 2 7.0 7.0 chydorus cf. sphaericus 20 89.5 19.7 acroperus sp. 3 22.2 12.2 alona glabra 13 28.6 6.7 alona quadrangularis type 12 42.9 10.7 alona cf. intermedia 1 5.6 alona manueli 6 40.0 9.6 alona ossiani 8 42.4 16.4 alona rustica 1 4.3 alonella excisa 3 4.3 2.5 alonella nana 3 13.6 7.1 alonella pulchella 4 6.1 3.7 anthalona verrucosa-group 5 14.3 5.9 camptocercus dadayi 6 30.8 13.7 disparalona dadayi 2 11.9 9.0 dunhevedia sp. 3 14.3 8.4 ephemeroporus sp. 2 10.5 9.0 euryalona sp. 11 25.0 11.3 graptoleberis testudinaria 3 3.6 2.5 kurzia sp. 4 13.3 7.2 leberis sp. 5 50.0 16.6 leydigia louisi louisi 10 34.4 7.8 leydigia acanthocercoides-group 6 51.7 15.9 leydigiopsis ornata 2 11.1 8.1 notalona sculpta 7 15.9 7.1 pleuroxus denticulatus 4 11.1 6.4 simocephalus sp. 2 22.4 19.0 nrr1 4 31.8 11.4 nrr2 1 25.0 nrr3 1 4.3 nrr4 1 4.3 n, occurrence; nrr1-4, not recognize remains type 1 to 4; nrr2, probably headshield of leberis sp.; nrr3, probably coronatella monacantha (sars 1901) group. no n c om me rci al us e o nly 158 m. wojewódka et al. fig. 4. results of multivariate statistical analysis applied to qualitative and quantitative results of species composition of subfossil cladocera, and environmental variables. a) cluster analysis showing species grouping according to reciprocal correlation level; the groups were distinguished for species that showed a high level of similarity >0.8; species composition of each group is presented in tab. 3; b) canonical correspondence analysis (cca) for species relative abundances and environmental variables; length and direction of arrows show respectively the strength and direction of a certain environmental variable. no n c om me rci al us e o nly subfossil cladocera of central america 159 1000 m asl. cca analysis showed the preference of the daphnia pulex-group for clear waters, such as lakes atitlán and ipala, that showed transparency of 4.2 m and 3.6 m, respectively. moreover, the contribution of daphniidae was negatively correlated with conductivity and was higher in waters with conductivity ≤100 µs cm–1. the results show the preference of daphniidae for highlands and mountain areas and clean, low-mineralized and nutrientpoor waters. similar environmental preferences by daphniidae were also suggest by hart (2004) and by mergeay et al. (2005) for african lakes. the occurrence of daphniidae species in lakes of central american highlands was also observed by peréz et al. (2013). bosminidae species were observed along the whole altitudinal gradient of the lake studied, from lowland, to highland and mountain lakes. cca analysis revealed a negative correlation of eubosmina (bosmina (e.) coregoni, bosmina (e.) longispina) and a positive correlation of bosmina longirostris with altitude. eubosmina was found in six lowland lakes, but also in two lakes located at >1000 m asl (atitlán and ipala), in lake yojoa (639 m asl) and in lake apastepeque (509 m asl). this suggests that even though eubosmina prefered lakes located at a lower altitude, this is probably not the main factor determining the distribution of this genus. on the other hand, eubosmina mainly occurred in waters with high conductivity. these species were dominant in highly mineralized (4520 µs cm–1) and brackish (salinity=2.5‰) waters of lakes salpetén, as well as in el rosario and lachuá, which showed conductivity values >900 µs cm–1. the presence of eubosmina, along with bosmina longirostris in lake salpetén, confirms the tolerance of bosminidae to waters with high mineral content and/or salinity (aladin, 1991). the obtained results also confirm the wide distribution of planktonic daphniidae and bosminidae in central america (elías-gutiérrez et al., 1999; korovchinsky, 2006). chydoridae was the most species rich group of cladocera in the sediment studied. chydorus cf. sphaericus was the most common species, as it was present in 20 out of the 29 surveyed lakes located both in lowland and mountain areas. the relative abundance of chydorus cf. sphaericus in the lake located at the highest altitude (lake magdalena, 2863 m asl) was almost 90%. cca analysis showed that none of the measured environmental factors had a significant effect on the distribution of the species in the region studied of central america. this confirmed the ubiquity of chydorus cf. sphaericus and its wide range of tolerance to environmental and ecological conditions (flössner, 2000; fryer, 1968; korhola and rautio, 2001; zawisza and szeroczyńska, 2011). the genus alona was mostly represented by alona glabra and the alona quadrangularis type. these species were found in 13 and 12 lakes, respectively, which were characterized by different conductivity, oxygenation, water visibility and altitude. the alona quadrangularis type was the unique littoral species found in lake salpetén, which may indicate its tolerance to waters with higher mineralization. the genus alona was also represented by alona manueli (sinev and zawisza, 2013) and alona ossiani, which, based on the cluster analysis, were grouped together, along with alonella nana (baird, 1843), alonella pulchella (herrick, 1884), the anthalona verrucosa-group and disparalona dadayi (birge, 1910). the multivariate analysis showed a negative correlation between group 3 and lake conductivity, which most likely resulted from the presence of species included in group 3 in lakes with conductivity ≤100 µs cm–1 and with medium conductivity (204-331 µs cm–1). lake sacnab (412 µs cm–1) and lake lachuá (906 µs/cm) are exceptions, and two species from group 3 (i.e., alona ossiani and alona manueli, respectively) were present. sinev (1998) showed that alona ossiani belongs to the alona affinis complex, which is considered an oligotrophic species in eurasia (kamenik et al., 2007; korhola and rautio, 2001). on the other hand, alona manueli (an endemic species of central america) and alonella pulchella were identified in oligotrophic lakes in mexico (cuna et al., 2014; sinev and zawisza, 2013). moreover, this studied provided the first record of alonella pulchella in the tropical region, as previously this species was recorded only in canada (korosi and smol, 2012). the obtained results seem to confirm the preference of these species for waters of low trophic status. the negative correlation of group 3 with the lake area may indicate the preference of these species for smaller lakes with well-developed littoral zone. in particular, alonella tab. 3. groups of species correlated with each other at the minimum level of 0.8 and distinguished based on the cluster analysis. group species composition 1 daphnia pulex-group, bosmina longirostris 2 bosmina (e.) coregoni, bosmina (e.) longispina, alona glabra, alona quadrangularis type, leberis sp. 3 alona manueli, alona ossiani, alonella nana, alonella pulchella, anthalona verrucosa-group, disparalona dadayi 4 euryalona sp., kurzia sp., leydigia acanthocercoides-group 5 acroperus sp., alona cf. intermedia, camptocercus dadayi, dunhevedia sp., pleuroxus denticulatus 6 alona rustica, simocephalus sp. no n c om me rci al us e o nly 160 m. wojewódka et al. nana, was present in only three small lakes with area ≤0.1 km2. this suggests that, similarly to what observe in europe, alonella nana may prefer small lakes (fryer, 1968; korosi and smol, 2012). furthermore, the presence of alona ossiani in mountain lakes may indicate a pronounced tolerance toward the harsher climatic conditions determined by altitude. the eurasian species alona affinis, is considered to tolerate lower temperatures (kamenik et al., 2007; locke and sprules, 2000) and the results of the present study indicated that allona ossiani shows similar characteristics in central america. leydigia louisi louisi and leydigia acanthocercoidesgroup played a significant role in lake el muchacho (almost 90%), which was characterized by a water transparency of only 0.4 m. moreover, these species were relatively abundant (>10%) in other lakes with water transparency below 0.9 m. the cca analysis confirmed the preference of these species for waters with lower secchi visibility. the presence of a species characteristic of central america such as leydigiopsis ornata, which is a typical benthic species (van damme and dumont, 2010), was recorded only in 2 shallow (<2.5 m water depth) and scarcely transparent lakes, which were characterized (secchi depth <1 m). graptoleberis testudinaria, which was recorded in the sediments of lakes atitlán, quexil and verde, preferred waters with higher secchi visibility (≥2.7 m), as showed also by the cca analysis. however, the results of the multivariate statistical analysis are highly controversial, because garptoleberis testudinaria is commonly considered to be associated with aquatic vegetation (fryer, 1968; rybak and błędzki, 2016). this suggests that other factors not included in our analysis control the distribution of this species. pleuroxus denticulatus (birge, 1879) remains were identified in the sediments from four lakes (chicabal, jucutuma, ticamaya, los negritos) and its maximum contribution to the cladocera communities did not exceed 12% (average=6.4%). the presence of only one pleuroxus species in the lakes of the study area and its local distribution confirms the rarity of this cladocera group in the tropical region (korovchinsky, 2006). canonical correspondence analysis outlined altitude as a key driver of the cladoceran fauna of in central america. other significant environmental factors affecting the cladocera distribution were water transparency and conductivity. among the environmental variables considered in this study, the less significant in affecting the waterflea population in the study area was the concentration of dissolved oxygen. conclusions this study showed a relatively typical distribution of cladocera in lakes of the neotropic region. in summary, our results indicate that: i) the most common cladocera species of the study area was chydorus cf. sphaericus, likely in relation to the ability of this species to adapt to different ecological conditions; ii) planktonic species of daphniidae and bosminidae were the most abundant; iii) daphniidae species were recorded primarily in highland lakes; iv) eubosmina and the alona quadrangularis type showed pronounced tolerance to waters with high mineralizatiovn level; e) alona ossiani, alona manueli, alonella nana and alonella pulchella were negatively correlated with water conductivity; vi) the presence of alonella nana was restricted to lakes with a smaller surface area; vii) alona ossiani was an important subdominant species in shallow lakes located at high altitudes; viii) among the environmental variables considered, altitude was a key driving factor for distribution of the cladocera fauna whereas dissolved water oxygenation was almost insignificant. these results provide basic information on cladocera communities in freshwater ecosystems of central america, a region where research carried out so far is insufficient to provide robust taxonomical and ecological information. acknowledgments the study was funded by the polish ministry of science (grant ncn 2014/13/b/st10/02534) and the german research foundation (dfg, schw 671/16-1). furthermore, scientific cooperation was supported by the polish german governments (mnisw-daad, 2016-2017). we would like to thank prof. a.y. sinev and prof. a.a. kotov for their help with identification of specimens. special thanks is due to to cuauhtémoc ruiz (instituto tecnológico de chetumal), ramón beltrán (centro interdisciplinario de ciencias marinas, mexico), and lisa heise (universidad autónoma de san luis potosí, mexico) for their excellent work on field. we also like to thank all people involved in this work: margarita caballero (instituto de geofísica, unam), alexis oliva and the team from the asociación de municipios del lago de yojoa y su área de influencia (amuprolago, honduras), maría reneé alvarez, margarita palmieri, eleonor de tott (universidad del valle de guatemala, guatemala), consejo nacional de áreas protegidas (conap, guatemala), néstor herrera and ministerio de medio ambiente (san salvador, el salvador). conacyt (mexico) provided fellowship (218604, 405326) for the third and fourth authors. references aladin nv, 1991. salinity tolerance and morphology of the osmoregulation organs in cladocera with special reference to cladocera from the aral sea. hydrobiologia 225:291-299. bigler c, heiri c, krskova r, lotter af, sturm m, 2006. disno n c om me rci al us e o nly subfossil cladocera of central america 161 tribution of diatoms, chironomids and cladocera in surface sediments of thirty mountain lakes in south-eastern switzerland. aquat. sci. 68:154-171. bjerring r, 2007. lake response to global change: nutrient and climate effects using cladoceran (crustacea) subfossils as proxies. phd thesis, national environmental research institute university of aarhus, denmark. campanelli mortari r, henry r, 2016. horizontal distribution of cladocera in a subtropical lake marginal to a river. j. paleolimnol. 75:109-120. cuna e, zawisza e, caballero m, ruiz-fernández ac, lozanogarcía s, alcocer j, 2014. environmental impacts of little ice age cooling in central mexico recorded in the sediments of a tropical alpine lake. j. paleolimnol. 51:1-14. dumont hj, 1994. on the diversity of the cladocera in the tropics. hydrobiologia 272:27-38. elías-gutiérrez m, ciros-pérez j, suárez-morales e, silva-briano m, 1999. the freshwater cladocera orders ctenopoda and anomopoda) of mexico, with comments on selected taxa. crustacean 72:171-186. elías-gutiérrez m, kotov aa, garfias-espejo t, 2006. cladocera (crustacea: ctenopoda: anomopoda) from southern mexico, belize and northern guatemala.zootaxa 119: 1-27. elías-gutiérrez m, suárez-morales e, gutiérrez-aguirre ma, silva-briano m, granados-ramírez jg, garfias-espejo t, 2008. [cladocera y copepoda de las aguas continentales de méxico].[book in spanish]. universidad nacional autónoma de méxico: 322 pp. feniova iy, razlutsky vi, palash al, 2011. temperature effects of interspecies competition between cladoceran species in experimental conditions. inland water. biol. 4:65-71. flössner d, 2000. [die haplopoda und cladocera (ohne bosminidae) mitteleuropas].[book in german]. backhuys publisher, leiden: 428 pp. frey dg, 1986. cladocera analysis, p. 667-692. in: b.e. berglund (ed.), handbook of holocene palaeoecology and palaeohydrology. j. wiley & sons, chichester. frolova l, nazarova l, pestryakova l, herzschuh u, 2014. subfossil cladocera from surface sediment in thermokarst lakes in northern siberia, russia, in relation to limnological and climatic variables, j. paleolimnol. 52:107-119. fryer g, 1968. evolution and adaptive radiation in the chydoridae (crustacea: cladocera): a study in comparative functional morphology and ecology. philos. t. r. soc. b 254:221-385. green j, 1995. altitudinal distribution of tropical planktonic cladocera. hydrobiologia 307:75-84. hammer o, harper dat, ryan pd, 2001. past paleontological statistics software package for education and data analysis. palaeontol. electron. 4:4-9. hart rc, 2004. cladoceran periodicity patterns in relation to selected environmental factors in two cascading warm-water reservoirs over a decade, hydrobiologia 526:99-117. hudec i, 2000. subgeneric differentiation within kurzia (crustacea: anomopoda: chydoridae) and a new species from central america. hydrobiologia 421:165-178. jeppesen e, leavitt p, de meester l, peder jensen j, 2001. functional ecology and paleolimnology: using cladocera remains to reconstruction anthropogenic impact. trends ecol. evol. 16:191-198. juggins s, 2005. new features in c2 version 1.4. university of newcastle. juggins s, 2007. user guide c2 software for ecological and paleoecological data analysis and visualisation user guide version 1.5. university of newcastle. kamenik c, szeroczyńska k, schmidt r, 2007. relationships among recent alpine cladocera remains and their environment: implications for climate-changes studies. hydrobiologia 594:33-46. karmalkar av, bradley rs, diaz hf, 2011. climate change in central america and mexico: regional climate model validation and climate change projections. clim. dyn. 37:605-629. kienast f, wetterich s, kuzmina s, schirrmeister l, andreev aa, tarasov p, nazarova l, kossler a, frolova l, kunitsky vv, 2011. paleontological records prove boreal woodland under dry inland climate at today’s arctic coast in beringia during the last interglacial. quat. sci. rev. 30:2134-2159. kotov aa, elías-gutiérrez m, guadalupe nieto m, 2003. leydigia louisi louisi jenkin, 1934 in the neotropics, l. louisi mexicana n. subsp. in central mexican highlands. hydrobiologia 510:239-255. kotov aa, 2009. a revision of leydigia kurz, 1875 (anomopoda, cladocera, branchiopoda), and subgeneric differentiation within the genus. magnolia press: 84 pp. korhola a, 1999. distribution patterns of cladocera in subarctic fennoscandian lakes and their potential in environmental reconstruction. ecography 22:357-373. korhola a, rautio m, 2001. cladocera and other branchiopod crustaceans, p. 5-41. in: j.p. smol, j.b. birks and w.m. last (eds.), tracking environmental change using lake sediments. zoological indicators. kluwer academic, dordrecht. korosi jb, smol jp, 2012. an illustrated guide to the identification of cladoceran subfossil from lake sediments in northeastern north america: part 2 the chydoridae, j. paleolimnol. 48:587-622. korovchinsky nm, 2006. the cladocera (crustacea: brachiopoda) as a relict group. zool. j. linn. soc. 147:109-124. locke a, sprules wg, 2000. effects of acidic ph and phytolankton on survival and condition of bosmina longirostris and daphnia pulex. hydrobiologia. 437:187-196. luoto t, sarmaja-korjonen k, nevalainen l, kauppila t, 2009. a 700 year record of temperature and nutrient changes in a small eutropied lake in southern finland. holocene 19:1063-1072. mergeay j, verschuren d, de meester l, 2005. daphnia species diversity in kenya, and a key to the identification of their ephippia. hydrobiologia 542:261-274. mirosław-grabowska j, zawisza e, 2014. late glaciale early holocene environmental changes in charzykowskie lake (northern poland) based on oxygen and carbon isotopes and cladocera data. quat. int. 328-329:156-166. nováková k, van hardenbroek m, van der knaap wo, 2013. response of subfossil cladocera in gerzensee (swiss plateau) to early late glacial environmental change. palaeogeogr. palaeoclimatol. palaeoecol. 391:84-89. padhye sm, kotov aa, dahanukar n, dumont hj, 2016. biogeography of the ‘water flea’ daphnia o.f. müller, (crustacea: branchiopoda: anomopoda) on the indian subcontinent. j. limnol. 75:1476. paterson mj, 1994. paleolimnological reconstruction of recent changes in assemblages of cladocera from acidified lakes no n c om me rci al us e o nly 162 m. wojewódka et al. in the adirondack mountains (new york). j. paleolimnol. 11:189-200. peréz l, lorenschat j, massaferro j, pailles ch, sylvestre f, hollwedel w, brandorff go, brenner m, islebe g, lozano ms, scharf b, schwalb a, 2013. bioindicators of climate and trophic state in lowland and highland aquatic ecosystems of the northern neotropics. rev. biol. trop. 61:603-644. rajapaksa r, fernando ch, 1987. redescription and assignment of alona globulosa daday, 1898 to a new genus notoalona and a description of notoalona freyi sp.nov. hydrobiologia 144:131-153. rey j, vasquez e, 1986. [contribution à la connaissance des cladocères néotropicaux: redescription de leydigiopsis ornata daday, 1905. (crustacea, cladocera)].[article in fench]. ann. limnol. 22:169-176. rybak ji, błędzik la, 2016. freshwater crustacean zooplankton of europe. springer: 918 pp. schmidt r, müller j, drescher-schneider r, krisai r, szeroczyńska k, barić a, 2000. changes in lake level and trophy at lake vrana, a large karstic lake on the islanad of cres (croatia), with respect to palaeoclimate and anthropogenic impacts during the last approx. 16,000 years. j. limnol. 59:113-130. shurin jb, 2000. dispersal limitation, invasion resistance, and the structure of pond zooplankton comunities. ecology 81:3074-3086. sinev ay, 1998. alona ossiani sp. n., a new species of the alona affinis complex from brazil, deriving from the collection of g.o. sars (anomopoda chydoridae). arthropoda sel. 7:103-110. sinev ay, 2001. redescripton of alona glabra sars, 1901, a south american species of the pulchella group (branchiopoda: anomopoda: chydoridae). arthropoda sel. 10:273-280. sinev a, zawisza e, 2013. comments on cladocerans of crater lakes of the nevado de toluca volcano (central mexico), with the description of a new species, alona manueli sp. nov. zootaxa 3647:390-400. sinev a y, 2015a. revision of the pulchella-group of alona s. lato leads to its translocation to ovalona van damme et dumont, 2008 (branchiopoda: anomopoda: chydoridae). zootaxa 4044:451-492. sinev ay, 2015b. morphology and phylogenetic position of three species of genus camptocercus baird, 1843 (cladocera: anomopoda: chydoridae). zootaxa 4040:169-186. sinev ay, dumont hg, 2016. revision of the costata-group of alona s. lato (cladocera: anomopoda: chydoridae) confirms its generic status. eur. j.taxon. 223:1-38. sinev ay, silva-briano m, 2012. cladocerans of genus alona baird, 1843 (cladocera: snomopoda: chydoridae) and related genera from aguascalientes state, mexico. zootaxa. 3569:1-24. sweetman jn, rühland km, smol jp, 2010. environmental and spatial factors influencing the distribution of cladocerans in lakes across the central canadian arctic treeline region. j. limnol. 69:76-87. szeroczyńska k,1991. impact of prehistoric settlements on the cladocera in the sediments of lakes suszek, błędowo, and skrzetuszewskie. hydrobiologia 225:105-114. szeroczyńska k, tatur a, weckstrom j, gąsiorowski m, noryśkiewicz a, sienkiewicz e, 2007. holocene environmental history in northwest finnish lapland reflected in the multi-proxy record of small subartic lake. j. paleolimn. 38:25-47. szeroczyńska k, zawisza e, 2011a. subfossil faunal and floral remains (cladocera, pediastrum) in two northern lobelia lakes in finland. knowl. manag. aquat. ect. 403:09. szeroczyńska k, zawisza e, 2011b. records of the 8200 cal bp cold event reflected in the compositionof subfossil cladocera in the sediments of three lakes in poland. quat. int. 233:185-193. szeroczyńska k, zawisza e, wojewódka m, 2015. initial time of two high altitude crater lakes (nevado de toluca, central mexico) recorded in subfossil cladocera. stud. quatern. 32:109-116. taylor ma, alfaro ej, 2005. climate of central america and the caribbean, p. 181-189. in: j.e. oliver (ed.), encyclopedia of world climatology. springer, dordrecht. ter braak cjf, 1986. canonical correspondence analysis: a new eigenvector technique for multivariate direct gradient analysis. ecology 67:1167-1179. van damme kv, dumont hj, 2010. cladocera of the lençóis maranhenses (ne brazil): faunal composition and a reappraisal of sars’ method. braz. j. biol. 70:755-779. van damme kv, sinev ay, dumont hj, 2011. separation of anthalona gen.n. from alona baird, 1843 (branchiopoda: cladocera: anomopoda): morphology and evolution of scraping stenothermic alonines. zootaxa 2875:1-64. zawiska i, słowiński m, correa-metrio a, obremska m, luoto t, nevalainen l, woszczyk m, milecka k, 2015. the response of a shallow lake and its catchment to late glacial climate changes a case study from eastern poland. catena 126:1-10. zawisza e, szeroczyńska k, 2011. cladocera species composition in lakes in the area of the hornsund fjord (southern spitsbergen)-preliminary results, knowl. manag. aquat. ec. 402:04. zawisza e, caballero m, ruiz-fernandez c, 2012. 500 years of ecological changes recorded in subfossil cladocera in a high-altitude tropical lake lago de la luna, central mexico. stud. quatern. 29:23-29. zawisza e, correa-metrio a, caballero m, lozano s, szeroczyńska k, 2014. paleoecology of tropical lake zirhauen (western mexico) recorded in cladocera remains. abstract book of the 8th shallow lakes conf., antalya, turkey. zawisza e, cuna e, caballero m, ruiz-fernandez ac, szeroczyńska k, woszczyk m, zawiska i, 2016. environmental changes during the last millennium recorded in subfossil cladocera, diatoms and sediment geochemistry from lake el sol (central mexico). geol. q. (in press). no n c om me rci al us e o nly layout 1 introduction the genus unruhdinium was erected in 2017 by gottschling et al. (2017) encompassing species previously included in the genus peridiniopsis (gottschling et al., 2017; moestrup and calado, 2018). it belongs to the monophyletic family of kryptoperidiniaceae (kretschmann et al., 2018 and references therein), which includes dinoflagellates hosting a tertiary diatom endosymbiont (i.e., dinotoms; imanian et al., 2011). the unruhdinium genus includes 7 freshwater species usually hosting cyclotellalike diatoms as endosymbionts. the morphological diagnostic traits of unruhdinium species are the reduced number of epithecal plates and the presence of five cingular plates (gottschling et al., 2017; moestrup and calado, 2018). all the species included in the genus were previously included in the genus peridiniopsis (gottschling et al., 2017; moestrup and calado, 2018). all unruhdinium species except u. jiulongensis (gu) gottschling and u. armebeense (ten-hage, k.p. da, couté) moestrup et calado, were reported as high biomass bloom formers in various lakes and reservoirs worldwide (rodriguez et al., 1999; liu et al., 2008; takano et al., 2008; zhang et al., 2011; zhang et al., 2014), including some italian alpine lakes (hansen and flaim, 2007). in mediterranean reservoirs, dinoflagellate blooms have been detected since the seventies and they have been mainly caused by ceratium hirundinella (mul̈ler) dujardin (lugliè et al., 2001; pérez-martìnez and sànchez-castillo, 2001; fadel et al., 2015; mariani et al., 2015) and gymnodinium uberrimum (g.j. allman) kofoid and swezy (fadda et al., 2016). in the last decade, intense blooms accompanied by reddish-brown water discolorations were recorded in the cedrino lake (sardinia, mediterranean region; padedda et al., 2017). on those occasions, the responsible dinoflagellate was first recognized as a peridinium species. a subsequent more accurate identification attributed the observed specimens to the genus peridiniopsis. in this study, the morphological and molecular identification of the species involved in the cedrino lake blooms were investigated on field samples and on cultured strains and thus determined as unruhdininium penardii (lemmermann) gottschling. the morphology and abundance of resting cysts from the sediments were also determined. furthermore, the field ecology of the species was explored. methods study area, sampling strategy and analyses cedrino lake is located in the central-eastern sardinia (western mediterranean, italy; fig. 1). its surface area is 1.5 km2 and the mean depth is 26.5 m. the lake is eutrophic and its waters are exploited for potable and agricultural uses (padedda et al., 2015, 2017). cyanobacteria, including potentially toxic species, dominate the phytoplankton composition especially in summer (messineo et al., 2009; article first detection of the bloom forming unruhdinium penardii (dinophyceae) in a mediterranean reservoir: insights on its ecology, morphology and genetics cecilia teodora satta,1* albert reñé,2 bachisio mario padedda,1 silvia pulina,1 giuseppina grazia lai,1 oriana soru,3 paola buscarinu,3 tomasa virdis,3 salvatore marceddu,4 antonella lugliè1 1department of architecture, design and urban planning, university of sassari, via piandanna 4, sassari, italy; 2institut de ciències del mar-csic, dpt. biologia marina i oceanografia, passeig marítim de la barceloneta 37−49, 08003, barcelona, spain; 3ente acque della sardegna (enas), via mameli 88, cagliari, italy; 4institut of sciences of food production, national research council (cnr), li punti, ss, italy abstract the freshwater genus unruhdinium includes dinoflagellates hosting a tertiary diatom endosymbiont. some of the species belonging to this genus form high-biomass blooms. in this study, data on the ecology, morphology and molecular identity of unruhdinium penardii were reported for the first time from a mediterranean reservoir (cedrino lake, sardinia, italy). the ecology of the species and its bloom events were examined along a multiannual series of data (2010-2017). cell morphology was investigated using field samples and six cultures established by cell isolation. a molecular identification of the six strains was performed. wild and cultured cells shared the same morphology, showing a prominent apical pore complex and two/three more or less prominent hypothecal spines as distinctive characters in light microscopy. molecularly, the six cultured strains corresponded to the same taxonomic entity with sequences only differing in a few polymorphic positions for the studied markers ssu rdna, lsu rdna, its and endosymbiont ssu rdna. all markers showed 99.5%−100% similarity with the available u. penardii sequences. seasonality of u. penardii revealed its preference for the colder semester (from december to june) with bloom events restricted to late winter/early spring months. three blooms resulting in reddish water discolorations were observed along the study period (2011, 2012 and 2017). glms revealed a significant role of water depth, temperature, and reactive phosphorous in determining the highest cell densities (>5 x 104 cells l–1). the results obtained contribute to the increase of field ecology knowledge on this species, demonstrating it is well established in the mediterranean area, and being able to produce recurrent high biomass blooms in the studied reservoir. no nco mm er cia l u se on ly c.t. satta et al.72 padedda et al., 2017). cedrino lake belongs to the italian network of long term ecological research (lter-italy; deims.org/9010f9db-3d6b-4253-9604-4e10f6714000). the regional sardinian agency ‘ente acque della sardegna’ (enas) is the manager of the cedrino lake since 2006. samplings were conducted from july 2010 to may 2018 at one station close to the deepest part of the lake (fig. 1) following different patterns (supplementary tab. 1). in fact, the data collection activities were carried out under various projects with different objectives. consequently, samplings were conducted monthly in 2010 and 2011 and bi-monthly from 2012 to 2018 (supplementary tab. 1). during the sampling period several interruptions occurred, mainly due to adverse weather conditions or sampling difficulties. water samples were collected from selected water depth layers using a niskin bottle. samples for phytoplankton and chlorophyll a (chla) analyses were collected from 0, 1, 2.5, 5, 7.5 and 10 m water depth layers. phytoplankton samples were immediately fixed in lugol’s iodine solution (1% final concentration) for the cell density estimate, determined following utermoḧl (1958) under an inverted microscope axiovert 25 (carl zeiss, oberkochen, germany) at 200x magnification. further non-fixed samples were taken for the observation of live cells, for cell culturing and for formalin-fixation. cell counts were made for each of the six depths from 2010 to 2013, whereas only one sample corresponding to the depth with the highest chla was counted from 2014 to 2018. chla was determined as described by goltermann et al. (1978). water temperature (temp), conductivity (con), dissolved oxygen (do) and ph were measured in situ with multi-parameter probes (ysi 6600 v2 and hydrolab ds5). transparency was measured with a secchi disk (sd). euphotic zone depth was calculated using sd measures (zeu: 2.5 times the sd depth; poikane, 2009). water samples for nutrient analyses were collected from fig. 1. geographical location of the cedrino lake (sardinia, italy) and sampling station placement. no nco mm er cia l u se on ly unruhdinium penardii in a mediterranean reservoir 73 each of the selected water depth layers (0, 1, 2.5, 5, 7.5, and 10 m) plus further depth layers along the entire water column (15, 20, and 30 m). concentrations of nutrients such as reactive (p-po4) and total (tp) phosphorus, ammonium (n-nh4), nitrate (n-no3), (n-no2) and total nitrogen (tn), were determined according to the methods of strickland and parsons (1972). total dissolved inorganic nitrogen (din) was calculated as the sum of nnh4, n-no3, and n-no2. surface sediment samples were taken with a grab at tab. 1. unruhdinium penardii cell densities in the samples corresponding to the highest chlorophyll a values at each data sampling in the photic zone and u. penardii cell densities and chlorophyll a values calculated as weighted average in the photic zone. data corresponding to u. penardii cell densities higher than 5 x 104 cells l–1 are reported in bold. u. pen depth chla u. pen wa chla wa disc (cells x 103 l–1) (m) (mg m–3) (cells x 103 l–1) (mg m–3) (m) 2010 jul 0 2.5 17.5 0 19.6 1.0 aug 0 1 18.5 0 17.3 1.5 sep 0 0 56.9 0 51.1 1.5 oct 0 0 14.6 0 11.5 1.2 dec 5.8 1 0.6 3.5 0.3 2.5 2011 jan 0 1 0.6 0 0.3 3.0 feb 0 7.5 1.1 2.1 0.7 1.1 mar 758.2 2.5 21.9 710.9 18.3 2.0 apr 5.9 5 24.1 16.6 18.8 1.7 may 0 0 0.9 1.6 0.6 3.0 jun 0 1 4.9 0 4.2 2.3 jul 0 1 8.2 0 7.0 1.7 aug 0 1 29.2 0 27.4 1.5 sep 0 0 58.8 2.6 50.9 1.0 dec 4.0 2.5 0.1 2.3 0.1 1.6 2012 feb 25,632 0 359.7 9380.9 151.92 0.7 apr 1.9 1 1.0 0.8 0.4 7 jun 10.1 0 101.8 10.5 99.0 0.6 aug 0 2.5 78.2 0 70.2 0.5 oct 0 1 29.1 0 24.7 1.8 dec 0 0 14.4 0.5 13.0 3.2 2013 feb 0 5 13.5 29.2 10.1 1.5 may 6.0 0 39.0 6.0 37.5 1 jul 0 2 14.6 0 9.9 1.4 sep 0 0 8.6 0 7.4 2.6 2014 jan 88 2.5 8.5 7.9 1.7 mar 32 2.5 29.7 27.0 1.7 may 0 0 9.4 9.1 2.5 2015 jan 3 1 1.9 1.6 1.7 mar 6 0 1.5 1.5 0.8 may 16 7.5 9.2 8.9 2.5 jul 0 7.5 5.4 3.7 2.6 oct 0 1 66.2 60.1 1.0 nov 0 2.5 6 5.3 3.2 2016 jan 0 2.5 1.7 1.6 3.0 mar 0 1 4.5 3.7 3.5 may 0 7.5 8.5 7.7 2.8 jul 0 5 9.9 6.9 2.0 sep 0 1 4.0 3.1 3.5 nov 0 7.5 7.6 5.8 2.6 2017 mar 4160 0 72.1 63.6 1.0 may 0 0 4.7 2.7 7.5 jul 0 2.5 30 27.8 1.8 nov 0 1 1.5 1.4 1.9 2018 jan 0 2.5 1.8 1.5 4.0 mar 0 7.5 6.4 4.7 4.4 may 32 0 34.2 25.2 1.2 u. pen, unruhdinium penardii; depth, sampling water depth; chla, maxima chlorophyll a; u. pen wa, u. penardii weighted density; chla wa, weighted chlorophyll a; disc, secchi disc. no nco mm er cia l u se on ly c.t. satta et al.74 the same station of water samples in may 2013. sediment samples were processed following the methods reported in satta et al. (2014), using filtered deionized water instead of filtered seawater. the resulting sample was counted in a 3-ml sedimentation chamber with a zeiss axiovert 10 inverted microscope at 400x magnification. cyst photographs were taken with a zeiss axiocam (carl zeiss, oberkochen, germany). cyst measurements were obtained from the photographs using the imagej software (1.47v; w. rasband, usa). cyst abundances were expressed as the number of cysts per gram dry weight of sediment. dry weight (dw) was obtained by drying wet sediment subsamples (2-3 cm3) at 105°c for 24 h. germination experiments were conducted in order to identify unruhdinium penardii cysts. single cysts and 1 ml of the sieved (100-10 µm; as described in satta et al., 2014) sample were placed in tissue culture multiplates filled with the abovementioned culture medium. the plates were incubated at 12±1°c, with 10:14 light:dark cycle, under an irradiance of 100 µmol m–2 s–1. plates were verified every 2−4 days to check cyst germination. strain origin and morphological and molecular characterization of vegetative cells six unruhdinium penardii strains (from upen1_uniss to upen6_uniss) were obtained from water samples collected in march 2017. the strains were cultured at a temperature of 12±1°c with an approximate illumination of 100 μmol photons m–2 s–1 and a photoperiod of 10:14 h light:dark (l:d). the cells were grown and maintained in l1-si medium (guillard and hargraves, 1993) prepared with filtered deionized water and supplemented by a soil extract (watanabe, 2005). the morphology of living and fixed cells from cultures and from the field was determined using axiovert 100 and axiovert 10 (carl zeiss) inverted microscopes equipped with epifluorescence and differential interference contrast optics. light microscopic (lm) examination of the thecal plate tabulation was performed on fixed cells (lugol’s iodine, 1% final concentration) stained with calcofluor white (fritz and triemer, 1985). chloroplast autofluorescence was examined in live cells. the shape and location of the nuclei were determined after staining 2% formalin-fixed cells for 10 min with 4’-6-diamidino-2-phenylindole (dapi). lm photographs were taken with a zeiss axiocam (carl zeiss; axiovert 10) and a spot flex digital camera spot imaging (sterling heights, mi, usa; axiovert 100). living and fixed cells from cultures and from the field were also observed under scanning electron microscopy (sem). cells were filtered into a 3.0–5.0 µm polycarbonate filter, and washed in distilled water for 15 min. a subsequent dehydration was carried out in a 25, 50, 75, 90, 95, and 100% ethanol series for ca. 10 min. the final step of 100% ethanol was repeated twice. the filters were critical-point dried in liquid co2 using a polaron jumbo (quorum technologies ltd, laughton, england) critical-point drying apparatus. the dried filters were then mounted on stubs, sputter coated with goldpalladium and viewed under a sem zeiss-evo (carl zeiss, oberkochen, germany). sem micrographs were presented on a black background using adobe photoshop 6.0 (adobe systems, san jose, ca, usa). cell size from cultured and wild fixed material was determined at 400x microscopic magnification using a calibrated eyepiece micrometer and from the lm and sem photographs using the imagej software (1.47 v). genomic dna of the six strains (from upen1_uniss to upen6_uniss) was extracted from ~15 ml of exponentially growing cultures. the cells were harvested by centrifugation at 3,000 rpm for 15 min. the pellet was transferred to a 2-ml microcentrifuge tube and centrifuged at 10,000 rpm for 5 min. total genomic dna was extracted from the final pellet using the ultraclean microbial dna isolation kit (mo bio laboratories, inc., carlsbad, ca, usa), following the manufacturer’s instructions. pcr amplifications followed standard protocols that are described in detail in satta et al. (2020). briefly, for ssu and lsu rdna, a first pcr was conducted to amplify the region of interest using euka-eukb (medlin et al., 1988) and d1r-d2c (scholin et al., 1994) primers respectively. the pcr products were used as a template for a subsequent nested pcr to obtain the dinoflagellate ssu rdna sequence, using dino18sf1-18scomr1 primers (lin et al., 2006; zhang and lin, 2005), the endosymbiont diatom ssu rdna sequence, using diaf-diar primers (zhang et al. 2011), and the dinoflagellate lsu rdna sequence using dinfi-dinri primers (logares et al., 2007). the its region was directly amplified using itsa-itsb primers (adachi et al., 1994). purification and sanger sequencing were conducted by an external service (genoscreen, lille, france). the lengths of the sequences obtained were approximately 1,750 bp (ssu rdna), 550 bp (lsu rdna), 520-650 bp (its) and 1,400 bp (endosymbiont ssu rdna). the genbank accession numbers are: mw194105 – mw19410 (ssu rdna), mw194984 – mw194988 (its) and mw195006 – mw195011 (lsu rdna) and mw217557 – mw217562 (endosymbiont ssu rdna). the sequences generated in this study were aligned with sequences obtained from genbank belonging to the genus unruhdinium, kryptoperidiniaceae representatives, and other dinoflagellate species used as outgroups. the dataset for each sequenced rdna region was aligned separately using mafft v. 7 (katoh et al., 2002) and manually verified using geneious v. r6 (biomatters ltd., new zealand). the ssu and lsu rdna datasets were trimmed using trimal (capella-gutiérrez et al., 2009). the its dataset, which is less conserved than previous regions, was trimmed using gblocks v0.91b (castresana 2000) using less no nco mm er cia l u se on ly unruhdinium penardii in a mediterranean reservoir 75 stringent parameters to avoid an excessive shorten of the alignment. for the endosymbiont sequences obtained, a dataset was constructed including available unruhdinium endosymbiont sequences and stephanodiscaceae diatoms sequences following čalasan et al. (2018) and aligned as previously described. the resulting alignments had 1695 (ssu rdna), 611 (its), 607 (lsu rdna) and 1432 (endosymbiont ssu rdna) positions. phylogenetic analyses were performed for each region using maximum likelihood (ml) and bayesian approaches. for the ml tree, the mpi version of raxml (randomized axelerated maximum likelihood; stamatakis 2014) was used. the best-fit model of nucleotide substitution was estimated using jmodeltest 2 (darriba et al., 2012). for all four alignments, the gtr model with a gamma distribution was selected, as it showed the best likelihood score (lsu and ssu rdna), or not significantly different than the best one (its, endosymbiont ssu rdna). the most likely tree was established from 1,000 searches. the ml bootstrap support (bs) was analyzed with 1,000 replicates. the bayesian analysis was carried out with mrbayes v. 3.2 (ronquist et al., 2012) using the same evolutionary model, four mcmc chains and 1,000,000 generations. the consensus tree was obtained from post burn-in trees. statistical support was evaluated by calculating the bootstrap values (%bs), and bayesian posterior probabilities (bpp) for all tree topologies. statistical analyses interactions between unruhdinium penardii data and the available environmental variables were analysed applying generalized linear models (glms). u. penardii observations in field samples were transformed to binary data (presence/absence) consequently the logit-link function for binomial distribution was applied (mccullagh and nelder, 1989). a first model was conducted on the basic presence/absence set of data. a second one was realized using an arbitrary threshold of u. penardii cell density set at 5 x 104 cells l–1. the fixed terms (predictor variables) of glms were the water collection depth layer (water depth), temp, cond, p-po4, tp, n-no3, n-nh4, nno2 and tn. analyses were performed in the statistical and programming software r version 3.3.3 using the package stats (r development core team, 2017). results morphological characterization of vegetative cells and cysts vegetative cells from field samples and cultures showed a conical epitheca and a hypotheca from conical to rounded (fig. 2 a,b). cell length and width ranged from 25 to 42.5 µm (mean of 31.4 µm ± 4.5, n=20) and from 23.8 to 35 µm (mean of 29.3 µm ± 3.9, n=20), respectively. the observations in lm showed the presence of two-three more or less prominent spines (fig. 2 a-c) and further shorter spines around the hypothecal plate borders (fig. 2 d,e). a distinct eyespot was observed in the sulcal area (fig. 2c). chloroplasts were numerous and discoid in shape (fig. 2f). nuclear staining with dapi showed the presence of two rather equal in size, rounded nuclei in equatorial position, one of which clearly showing condensed chromosomes (dinokarion) (fig. 2g). the plate formula was po, x, 4’, 0a, 6’’, 5c, 5s, 5’’’ and 2’’’’ (fig. 3 a-g). theca ornamentation consisted in numerous scattered pores (fig. 3 a-d). moreover, numerous cells also showed an irregular vermicular ornamentation (fig. 3 a,d,e,g). the apical pore complex (apc) was fairly in the centre of epitheca, showing two protruded rims on its sides (fig. 3 a,b,d). the pore plate (po) and the canal plate (x) formed the apc (fig. 3b). po was comma-shaped, 1.6±0.1 µm long (n=8), with the upper edge oriented towards the right side of the cell. the x canal was rectangular and 1.7±0.2 µm long (n=8). the apical plate series was symmetrical, without intercalary plate (fig. 3b). the first apical plate (1’) was pentagonal and contacted to apc, the fifth cingular plate (c5) and the sulcal anterior plate (sa) (fig. 3a). the 3’ plate was six sided (fig. 3b). the precingular plates (from 1’’ to 6’’) were four or five-sided (fig. 3 b,c). the cingulum was descending, displaced 1 time its width and provided of thorny rims more or less protruded (fig. 3 a, d). there were 5 cingular plates, of which c1 was the narrowest (fig. 3a). five plates formed the sulcus (fig. 3e): the sulcal anterior plate (sa), the posterior sulcal plate (sp), the right sulcal plate (sr), the left sulcal plate (sl) and the median sulcal plate (sm). sa was the smallest, whereas sp the largest (fig. 3a). sr was elongated and supported a sulcal list. sl and sm were slightly smaller than sr and they were both almost hidden by the sr list (fig. 3e). a ventral pore was visible at the base of sa (fig. 3a). the two antapical plates (1’’’’ and 2’’’’) were rather centred with respect to the five postcingular plates (1’’’−5’’’) and the sp (fig. 3g). the antapical plates supported a series of spines (fig. 3g). the most prominent spines ranging from 0.7 to 1.8 µm in length (mean = 1.2±0.3 μm, n=20) were located along the edges adjacent to the sulcal area (fig. 3 a,e). this type of ornamentation, although less prominent, was also present along the postcingular plates (fig. 3 a, e-g). cysts were oval in shape, 29.6±2.2 μm long and 28.7±2.2 μm wide (n=11). the cyst content was grainy with yellowish to reddish accumulation bodies (fig. 4 a,b). most of the cysts maintained the external theca (fig. 4a). cysts with theca were more abundant in the sediments no nco mm er cia l u se on ly c.t. satta et al.76 fig. 2. morphology of unruhdinium penardii cells in light and epifluorescence microscopy from the cedrino lake (sardinia, italy). a,b) ventral and dorsal view of alive cells (a-b); c) eyespot localization; d,e) hypothecal spines arrangement; f) shape of chloroplasts; g) position of the nuclei, showing the dinokarion (n) and the diatom nucleus (n). scale bars: 10 μm. fig. 3. scanning electron micrographs showing the tabulation of unruhdinium penardii from the cedrino lake (sardinia, italy): ventral view (a), apical view showing the apc structure (b), dorsal view (c), latero-apical view (d), sulcal plates arrangement (e), dorsal view showing the dorsal post cingular plates (f), and antapical view (g). scale bars represent 10 μm for a-d and f-g. scale bar of e represents 2 µm. no nco mm er cia l u se on ly unruhdinium penardii in a mediterranean reservoir 77 showing a cyst density of 2807 cysts gr–1 dw, whereas the other smooth, oval cysts were detected at a cyst density of 792 cysts gr–1 dw. the incubation of sieved sediment samples produced unruhdinium penardii vegetative cells. the single-cyst germination experiments were unsuccessful. molecular characterization all six cultured strains produced almost identical sequences for the three rdna regions analysed (99.8% pairwise identity for ssu rdna sequences, 100% for lsu rdna and 99.8% for its rdna, respectively). most differences were found related to polymorphic positions. for ssu rdna phylogeny (fig. 5), a cluster including all kryptoperidiniaceae members, i.e. galeidinium, kryptoperidinium, durinskia, blixaea and unruhdinium genera, was obtained with moderate/high support (88%/1). sequences representing unruhdinium species formed a highly supported clade (99%/1). sequences from this study clustered with u. penardii sequences available from china (hm596543) and japan (ab353771), even though under moderate support (73%/0.96). the sequence available for u. niei (liu et hu) gottschling was at the base of this cluster, and the sequences of u. minimum (zhang, liu et hu) gottschling, u. cf. kevei and u. jiulongensis clustered independently. the same phylogenetic relationships were obtained for its (supplementary fig. 1) and lsu rdna (supplementary fig. 2), with all unruhdinium sequences forming a wellsupported clade (97%/1 for its and 94%/1 for lsu rdna). the sequences from this study clustered with u. penardii sequences available from china (100%/1 for its, 92%/for lsu rdna). the sequence available for u. penardii var. robustum (zhang, liu et hu) gottschling clustered at the base of the previous clade (85%/0.98) in the its phylogeny, and u. niei sequences forming a sister branch (-/0.93). for the lsu rdna phylogeny, the u. penardii var. robustum sequence formed a sister branch with u. niei sequences (-/0.99) clustering with u. penardii sequences (78%/0.99). in both phylogenies, sequences of u. jiulongensis and u. minimum clustered more distantly. for the endosymbiont ssu rdna, all sequences were also almost identical (99.7% pairwise identity) and differences corresponded to polymorphic positions. the constructed phylogeny (supplementary fig. 3) showed the sequences obtained were included in a well-supported cluster (95%/1) containing all the endosymbiont sequences available for unruhdinium representatives, together with the diatoms discostella/cyclotella. however, the genetic distance between members of the clade was low, resulting in several polytomies in internal branches and the relationships within this cluster were not well resolved. the sequences from this study clustered with sequences corresponding to u. cf. kevei and u. jiulongensis endosymbionts and two discostella/cyclotella sequences, but showing no statistical support (72%/-). instead, available u. penardii endosymbiont sequences clustered together with discostella nipponica in a sister clade, but again showing moderate statistical support (89%/-). polymorphic positions detected in the generated sequences mostly coincided with the differences between sequences available for u. penardii and u. cf. kevei endosymbionts, resulting in the observed topology. unruhdinium penardii dynamic and relationship with the environmental conditions along the study period unruhdinium penardii was detected in a wide seasonal time-frame (from december to june), embracing winter-cold and spring-mild months. on the contrary, its blooms were limited to a restricted seasonal period (february-march; tab. 1). u. penardii specimens were prevalently recorded in more surface samples (from 0 to 2.5 m). the three most relevant bloom events accompanied by reddish water discolorations were detected, in march 2011, february 2012 and march 2017 (tab. 1). the highest cell density was recorded in february 2012 (25.6 x 106 cells l–1). during the bloom events, chla values were among the highest found in the sampling period (maximum of 360 mg m–3 in the surface sample, in february 2012; tab. 1). the three bloom events corresponded to late winter-early spring environmental conditions in cedrino lake (supplementary tab. 2). temp values ranged from 11.1 °c to 15 °c, cond values from 313 µs cm–1 to 349 µs cm–1 and the lake was in mixing (supplementary tab. 2). nutrient concentrations revealed values among the lower of the entire dataset for p-po4 and intermediate for tp and din (supplementary tab. 2). the first glm regression revealed that temp and cond had significant effects on the presence of unruhdinium penardii along the study period (tab. 2). the second glm regression showed a significant effect fig. 4. morphology of unruhdinium penardii cysts in light microscopy from the cedrino lake (sardinia, italy): cyst with theca (a) and without theca (b). scale bars: 10 µm. no nco mm er cia l u se on ly c.t. satta et al.78 of water depth, temp and p-po4 on the presence of the species at the highest cell densities (tab. 2). discussion data collected in this study contribute to increase the information on the ecology, morphology and phylogeny of unruhdinium penardii. this work signals for the first time in a mediterranean reservoir, its well-established presence and its bloom forming behaviour. the collection of new morphological and molecular data is particularly of interest for freshwater dinoflagellates, many of which were described before electron microscopy and molecular genetic methods became available (annenkova, 2013). in addition, deepening the knowledge on the ecology of bloom-forming dinoflagellate species can be useful to manage events that may adversely affect water quality in mediterranean reservoirs, especially for their strategic role as drinking water resources in this semi-arid climate region (mariani et al., 2015). morphology and phylogeny in recent years, the combination of morphological and genetic studies allowed to identify and delineate the main characteristics of the peculiar freshwater dinoflagellates now included in the genus unruhdinium, formerly fig. 5. maximum likelihood phylogenetic tree inferred from the ssu rdna sequences, including representatives of the family kryptoperidiniaceae and other sequences used as outgroups. unruhdinium penardii sequences obtained in this study are indicated in bold. the bootstrap values (bp) and the bayesian posterior probabilities (bpp) are provided at each node (% bs/bpp). only bs and bpp values >70% and >0.90 are shown. no nco mm er cia l u se on ly unruhdinium penardii in a mediterranean reservoir 79 ascribed to peridiniopsis (gottschling et al., 2017). these new data derived from the description of new species (zhang et al., 2014; you et al., 2015) and from insights on known species (takano et al., 2008; zhang et al., 2011) underlining the growing interest and study effort on freshwater dinoflagellates belonging to the kryptoperidiniaceae family. following the collected morphological information, other dinoflagellates previously ascribed to the genus peridiniopsis were also recently included in the genus unruhdinium (u. armebeense and u. durandii (rodriguez, couté, tenhage et mascarell) moestrup et calado; moestrup and calado, 2018). morphological and molecular data obtained in this study allowed the determination of the species u. penardii from the field samples and from cultures of cedrino lake. consequently, data reported in this study are the first obtained on u. penardii from the mediterranean area. u. penardii from cedrino lake showed a prominent apical pore complex, clearly visible also in lm, four apical plates, no epithecal intercalary plates, five cingular plates and two or three antapical spines. this thecal plate pattern coincided with the recent published observations (hansen and flaim, 2007; takano et al., 2008; zhang et al., 2011, 2013). the presence of additional shorter spines along the antapical plate borders was not reported prior to this study, although hansen and flaim (2007) described the presence of small antapical spines in their material. spine lengths in the cedrino lake material were variable (maximum length of 1.8 µm) and spines were sometimes difficult to be observed in lm. the presence of small antapical spines differentiated the specimens from cedrino lake from those of u. penardii var robustum, whose morphological description indicated a large number of rather prominent spines (mean length: 2.8 µm; zhang et al., 2011). the other two characteristics already found in u. penardii and also observed in the cells from cedrino lake were the presence of an eyespot and a second eukaryotic nucleus. the morphological features of u. penardii (especially cell shape, size and presence of antapical spines) appeared to be quite specific in respect to the other unruhdinium species. on the contrary, these features were shared with u. durandii morphology. the main differences between the two species appeared to be the shape of the cell (ovoid in u. penardii and rhomboid in u. durandii) and the ornamentation of the plates (clearly reticulated in u. durandii). however, hansen and flaim (2007) reported a certain variability in u. penardii cell shape (from ovoid to rhomboid) and plate ornamentation (from smooth to slightly reticulated surface) suggesting u. durandii as junior synonymous of u. penardii. the cells from cedrino lake showed certain variability in the shape of the cells and a pseudo-reticulated ornamentation of the thecal plates. in addition, the distribution pattern of u. durandii spines and their lengths (maximum of 2 µm; rodriguez et al., 1999) matched with the observations obtained in this study. further data from the type locality of u. durandii are certainly needed to completely understand and describe the relationship between these two species. u. penardii cysts from cedrino lake sediments were oval in shape resembling those described by sako et al. (1987) from cultures. however, the most abundant cysts retained the theca, therefore closely resembling vegetative cells. similar benthic stages were described for blixaea quinquecornis (abé) gottschling (as peridinium quinquecorne; satta et al., 2010), bysmatrum subsalsum (ostenfeld) faust et steidinger (anglès et al., 2017) and from other both planktonic and benthic dinoflagellate species (bravo and figueroa, 2014 and references therein). the role and the origin of these ‘thecate cysts’ have yet to be investigated. unruhdinium penardii sequences from cedrino lake were identical or really close (>99.5% similarity) to those from chinese and japanese isolates. while some of the genetic markers used in this study, i.e., ssu rdna, correspond to conserved regions, lsu rdna and its regions are considered as more variable, and have been commonly used to determine intraspecific variability and speciation processes in dinoflagellates (penna et al., 2008; lin et al., 2020). the its sequences obtained for u. penardii were 100% identical to those from geographically distant locations (china and japan), suggesting a low intraspecific genetic variability, even tab. 2. results of the generalized linear models (glms) applied on selected environmental variables and on 1) the absence/presence data of unruhdinium penardii and 2) the absence/presence data using an arbitrary threshold of u. penardii cell density set at 5 x 104 cells l–1. significant effects are shown in bold. 1 glm absence/presence estimate standard error z value pr (>z) (intercept) 10.68 2.73 3.91 <0.0001 depth -0.06 0.06 -1.09 0.28 temp -0.37 0.09 -4.35 <0.0001 cond -0.01 0.01 -2.09 <0.05 p-po4 -0.03 0.02 -1.69 0.09 2 glm absence/presence (threshold: 5 x 104 cells l–1) estimate standard error z value pr (>z) (intercept) 81.67 34.86 2.34 <0.05 depth -0.80 0.39 -2.04 <0.05 temp -3.98 1.82 -2.17 <0.05 cond -0.08 0.05 -1.67 0.09 p-po4 -0.54 0.23 -2.31 <0.05 depth, sampling water depth; temp, water temperature; cond, conductivity; p-po4, reactive phosphorus. no nco mm er cia l u se on ly c.t. satta et al.80 though inhabiting isolated environments also very far from each other. the sequences obtained were clearly differentiated from those available for u. penardii var. robustum, in agreement with morphological differences described. in all rdna regions, the species showed the closest phylogenetic relationship with u. niei, rather than the other unruhdinium species, even though they did not show apparent morphological similarities supporting it. regarding the phylogenetic relationships of kryptoperidiniaceae, all sequences formed a wellsupported cluster for all regions, as also observed for the other genera of the family. however, the phylogenetic relationship among the different genera was not resolved. the information retrieved from the diatom endosymbiont ssu rdna sequences was not conclusive, even though it was in agreement with previous knowledge on the identity of the diatom endosymbiont of unruhdinium. the low genetic distance found for the different unruhdinium endosymbiont species and the presence of several polymorphic positions impeded a clear discrimination between u. penardii, u. cf. kevei and u. jiulongensis endosymbionts and to determine their phylogenetic relationships. ecology reservoirs are man-made ecosystems assuring important goods and services to local human populations (e.g., they are sources for drinking and irrigation water, energy production, aquaculture exploitation, flood management, touristic and leisure activities; padedda et al., 2017). their role as the main source of water supply for human populations was reported to be even more important in semi-arid regions, such as those of the mediterranean climate (marcé and armengol, 2010). eutrophication was indicated as the main issue impacting on reservoirs (smith et al., 1999) also in the mediterranean region (marchetto et al., 2009). among the detrimental effects of eutrophication, the occurrence of harmful algal blooms represented a consistent threat, especially for drinking and irrigation uses when the causative organisms were toxic (codd 2000). generally, the organisms involved in this kind of events belonged to cyanobacteria (paerl and otten, 2013; meriluoto et al., 2017). however, dinoflagellates were also frequently reported to cause blooms in reservoir ecosystems causing water discoloration and increase in water turbidity (e.g., ceratium hirundinella, lugliè et al., 2001; pérezmartìnez and sànchez-castillo, 2001; fadel et al., 2015; mariani et al., 2015). some species belonging to the unruhdinium genus, including u. penardii, were reported as high biomass bloom-forming species often causing dark or reddish water discolorations (javornický 1972; leitão et al., 2001; rodriguez et al., 1999; hansen and flaim, 2007; liu et al., 2008; you et al., 2015). the same typology of events was observed in cedrino lake during u. penardii blooms. however, ecological data describing the adverse events are today limited. field observations collected in this study allowed defining the ecological conditions during the presence of the species and at the highest densities. the species was detected along a fairly large temporal period (about coinciding with the first semester of the year) but it was never detected in summer. blooms exhibited strong seasonality restricted to the late winter and early spring months (february and march). glms revealed the significant role of temperature on u. penardii presence and blooms in the cedrino lake, underlining a preference of this species for colder months. the same seasonal behaviour was reported for u. niei in lake donghu (liu et al., 2008). in the cedrino lake, u. penardii produced blooms in a temperature range between 11.1°c and 15°c, whereas the same species was reported to bloom at lower temperature in the black sea (terenko 2017). u. durandii in the voiglans reservoir (rodriguez et al., 1999) showed the same temperature bloom range of u. penardii in the cedrino lake. investigating the role of nutrients is a challenge and must takes in consideration the possible mixotrophy of dinoflagellates in freshwater ecosystems (e.g., carrias et al., 2001; tardio et al., 2003; pålsson and granéli, 2004). anyway, the cedrino lake showed a high nutrient availability being classified from eutrophic to hypereutrophic (padedda et al., 2015, 2017). data from the present study revealed a significant association of u. penardii with low phosphate values, especially during blooms. kawabata and hirano (1995) observed a reduction of the cellular phosphorus content during the growth of u. penardii in the ishitegawa reservoir, hypothesizing a use of the stored phosphorus during the population growth phase. maximum cell densities of u. penardii in the cedrino lake could be indirectly explained by the ability of dinoflagellates to use organic sources of phosphorus (e.g., dop) as already revealed for u. niei (reported as peridiniopsis sp.; cao et al., 2019). these species, in this way, could acquire an advantage over the other phytoplankton species in low phosphate conditions. in addition, a strong significant correlation between unruhdinium cell density and the phosphatase alkaline activity was demonstrated from the three gorges reservoirs (reported as peridiniopsis sp.; wu et al., 2018). the same study revealed the ability of the species to migrate in deeper layers during the night in relation to the availability of dissolved organic phosphorus (wu et al., 2018). diel vertical migrations of u. niei were also in relationship to light radiation (xu et al., 2010). in fact, maximum cell densities of u. niei were recorded during the day and at the surface water layers (xu et al., 2010). similarly, u. penardii at the highest cell densities in the cedrino lake showed a significant inverse relationship no nco mm er cia l u se on ly unruhdinium penardii in a mediterranean reservoir 81 with water depth as revealed by glm. taking into account these observations, further investigations are required to completely understand the dynamics of the u. penardii blooms in the cedrino lake and to assess the role of the different environmental variables. in any case, data collected in this study provide a useful base of knowledge for further insights on this harmful species in the mediterranean area. conclusions in the scenario of global climate change underway, the mediterranean area was indicated as one of the most vulnerable (erol and randhir, 2012) and the aquatic resources of this area are reported to be very sensitive to these changes (garcía-ruiz et al., 2011). in particular, further water restrictions are expected. the impact of high biomass bloom forming dinoflagellates (as u. penardii) could be deleterious in drinking water plants. numerous species could cause unpleasant taste and odour of water, but also accelerate clogging of filter systems in drinkingwater treatment. they could also break through these filters with the consequence of elevating the dissolved organic carbon (doc) concentrations of the purified water and thus enhancing microbial growth (niesel et al., 2007). in this context, acquiring ecological data on potentially harmful species is mandatory in order to be able to manage adverse events maximizing the sustainable use of resources. acknowledgments the authors thank dr. bastianina manca and dr. pasqualina farina (university of sassari) for the nutrient analysis. the activities of prof. antonella lugliè and dr. bachisio mario padedda were supported by the research fund of the university of sassari (fondo di ateneo per la ricerca 2019). corresponding author: ctsatta@uniss.it key words: peridiniopsis; cultures; temperature; cysts; endosymbiont. received: 13 november 2020. accepted: 10 december 2020. this work is licensed under a creative commons attribution noncommercial 4.0 license (cc by-nc 4.0). ©copyright: the author(s), 2020 licensee pagepress, italy advances in oceanography and limnology, 2020; 11:9500 doi: 10.4081/aiol.2020.9500 references adachi m, sako y, ishida y, 1994. restriction fragment length polymorphism of ribosomal dna internal transcribed spacer and 5.8sregions in japanese alexandrium species (dinophyceae). j. phycol. 30:857-863. anglès s, reñé a, garcés e, lugliè a, sechi n, camp j, satta ct, 2017. morphological and molecular characterization of bysmatrum subsalsum (dinophyceae) from the western mediterranean sea reveals the existence of cryptic species. j. phycol. 53:883-847. annenkova nv, 2013. phylogenetic relations of the dinoflagellate gymnodinium baicalense from lake baikal. centr. eur. j. biol. 8:366-373. bravo i, figueroa ri, 2014. towards an ecological understanding of dinoflagellate cyst functions. microorganisms 2:11-32. čalasan až, kretschmann j, gottschling m, 2018. absence of co-phylogeny indicates repeated diatom capture in dinophytes hosting a tertiary endosymbiont. org. divers. evol. 18:29-38. cao x, wan l, xiao j, chen x, zhou y, wang z, song c, 2019. environmental effects by introducing potamogeton crispus to recover a eutrophic lake. sci. total environ. 621:360-367. capella-gutiérrez s, silla-martínez jm, gabaldón t, 2009. trimal: a tool for automated alignment trimming in largescale phylogenetic analyses. bioinformatics 25:1972-1973. carrias jf, thouvenot a, amblard c, sime-ngando t, 2001. dynamics and growth estimate of planktonic protists during early spring in lake pavin, france. aquat microb. ecol. 24:163-174. castresana j, 2000. selection of conserved blocks from multiple alignments for their use in phylogenetic analysis. mol. biol. evol. 17:540-552. codd ga, 2000. cyanobacterial toxins, the perception of water quality, and the prioritisation of eutrophication control. ecol. eng. 16:51-60. darriba d, taboada g, doallo r, posada d, 2012. jmodeltest 2: more models, new heuristics and parallel computing. nat methods 9:772. fadda a, manca m, camin f, ziller l, buscarinu p, mariani ma, padedda bm, sechi n, virdis t, luglie ̀a, 2016. study on the suspended particulate matter of a mediterranean artificial lake (sos canales lake) using stable isotope analysis of carbon and nitrogen. ann. limnol. int. j. limnol. 52:401-412. fadel a, atoui a, lemaire bj, vinco̧n-leite b, slim k, 2015. environmental factors associated with phytoplankton succession in a mediterranean reservoir with a highly fluctuating water level. environ. monit. assess. 187:633. fritz l, triemer re, 1985. a rapid simple technique utilizing calcofluor white m2r for the visualization of dinoflagellate thecal plates. j. phycol. 21:662-664. garciá-ruiz jm, loṕez-moreno ji, vicente-serrano sm, lasantamartińez t, begueriá s, 2011. mediterranean water resources in a global change scenario. earth. sci. rev. 105:121-139. goltermann hl, clymo rs, ohnstad mam, 1978. method for physical and chemical analysis of fresh waters. i.b.p. n. 8. blackwell scientific publications, oxford: 214 pp. gottschling m, čalasan až, kretschmann j, gu h,2017. two new generic names for dinophytes harbouring a diatom as an endosymbiont, blixaea and unruhdinium no nco mm er cia l u se on ly c.t. satta et al.82 (kryptoperidiniaceae, peridiniales). phytotaxa 306: 296-300. guillard rrl, hargraves pe, 1993. stichochrysis immobilis is a diatom, not a chrysophyte. phycologia 32:234-236. hansen g, flaim g, 2007. dinoflagellates of the trentino province, italy. j limnol. 66:107-141. imanian b, pombert jf, keeling pj, 2011. the complete plastid genomes of the two ‘dinotoms’ durinskia baltica and kryptoperidinium foliaceum. plos one 5:e10711. javornický p, 1972. peridinium penardii (lemm.) lemm. f. californicum, forma nova. j. phycol. 7:303-306. kretschmann j, čalasan až, gottschling m, 2018. molecular phylogenetics of dinophytes harbouring diatoms as endosymbionts (kryptoperidiniaceae, peridiniales), with evolutionary interpretations and a focus on the identity of durinskia oculata from prague. mol. phylogenetics evol. 118:392-402. leitão m, ten-hage l, mascarell g, coute ́a, 2001. peridiniopsis corillionii sp. nova (dinophyta), une nouvelle dinophycée d’eau douce de france responsible de marées rouges en rivier̀e. algol. stud. 102:1-15. lin s, zhang h, hou y, miranda l, bhattacharya d, 2006. development of a dinoflagellate-oriented pcr primer set leads to detection of picoplanktonic dinoflagellates from long island sound. appl. environ. microbiol. 72:5626 -5630. lin s, hu z, deng y, shang l, gobler cj, tang yz, 2020. an assessment on the intrapopulational and intraindividual genetic diversity in lsu rdna in the harmful algal bloomsforming dinoflagellate margalefidinium (= cochlodinium) fulvescens based on clonal cultures and bloom samples from jiaozhou bay, china. harmful algae 96:101821. liu gx, pei gf, hu zy, 2008. peridiniopsis niei sp. nov. (dinophyceae), a new species of freshwater red tide dinoflagellates from china. nova hedwigia 87:487-499. logares r, shalchian-tabrizi k, boltovskoy a, rengefors k, 2007. extensive dinoflagellate phylogenies indicate infrequent marine–freshwater transitions. mol. phylogenetics evol. 45:887-903. marcé r, armengol j, 2010. water quality in reservoirs under a changing climate, p. 73−94. in: s. sabater and d. barceló (eds.), water scarcity in the mediterranean. springer, amsterdam. marchetto a, padedda bm, mariani ma, lugliè a, sechi n, 2009. a numerical index to evaluate the phytoplankton response to changes in nutrient levels in deep mediterranean reservoirs. j limnol. 68:106-121. mccullagh p, nelder ja, 1989. generalized linear models. chapman and hall, london: 512 pp. meriluoto j, blaha l, bojadzija g, bormans m, brient l, codd ga, drobac d, faassen ej, fastner j, hiskia a, ibelings bw, kaloudis t, kokocinski m, kurmayer r, pantelić d, quesada a, salmaso n, tokodi n, triantis tm, visser pm, svirčev z, 2017. toxic cyanobacteria and cyanotoxins in european waters – recent progress achieved through the cyanocost action and challenges for further research. aiol 8:161-178. medlin l, elwood hj, stickel s, sogin ml, 1988. the characterization of enzymatically amplified eukaryotic 16slike rrna-coding regions. gene 71:491-499. messineo v, bogialli s, melchiorre s, sechi n, lugliè a, casiddu p, mariani ma, padedda bm, di corcia a, mazza r, carloni e, bruno m, 2009. cyanotoxins occurrence in italian freshwaters. limnologica 39:95-106. moestrup ø, calado aj, 2018. süßwasserflora von mitteleuropa, bd. 6 freshwater flora of central europe, vol. 6: dinophyceae. springer spektrum, berlin: 561 pp. niesel v, hehn e, sudbrack r, willmitzer h, chorus i, 2007. the occurrence of the dynophyte species gymnodinium uberrimum and peridinium willei in german reservoirs. j. plankton res., 29:347-357. padedda bm, sechi n, lai gg, mariani ma, pulina s, satta ct, bazzoni am, virdis t, buscarinu p, lugliè a, 2015. a fastresponse methodological approach to assessing and managing nutrient loads in eutrophic mediterranean reservoirs. ecol. eng. 85:47-55. padedda bm, sechi n, lai gg, mariani ma, pulina s, sarria m, satta ct, virdis t, buscarinu p, lugliè a, 2017. consequences of eutrophication in the management of water resources in mediterranean reservoirs: a case study of lake cedrino (sardinia, italy). glob. ecol. conserv. 12:21-35. paerl hw, otten tg, 2013. harmful cyanobacterial blooms: causes, consequences, and controls. microb. ecol. 65: 995-1010. pålsson c, granéli w, 2004. nutrient limitation of autotrophic and mixotrophic phytoplankton in a temperate and tropical humic lake gradient. j. plankton res. 26:1005-1014. penna a, fraga s, masó m, giacobbe mg, bravo i, garcés e, vila m, bertozzini e, andreoni f, lugliè a, vernesi c, 2008. phylogenetic relationships among the mediterranean alexandrium (dinophyceae) species based on sequences of 5.8s gene and internal transcript spacers of the rrna operon. eur. j. phycol. 43:163-178. pérez-martinez c, sánchez-castillo p, 2002. winter dominance of ceratium hirundinella in a southern north-temperate reservoir. j. plankton res. 24:89-96. poikane s, 2009. water framework directive intercalibration technical report. part 2:llakes. jrc scientific and technical reports. european commission, luxembourg: 176 pp. rodriguez s, couté a, ten-hage l, mascarell g, 1999. peridiniopsis durandi sp. nova (dinophyta), une nouvelle dinophyceé d’eau douce responsible de mareés rouges. algol. stud. 95:15-29. ronquist f, teslenko m, van der mark p, ayres dl, darling a, höhna s, larget b, liu l, suchard ma, huelsenbeck jp, 2012. mrbayes 3.2: efficient bayesian phylogenetic inference and model choice across a large model space. syst. biol. 61:539-542. satta ct, anglès s, garcés e, sechi n, pulina s, padedda bm, stacca d, lugliè a, 2014. dinoflagellate cyst assemblages in surface sediments from three shallow mediterranean lagoons (sardinia, north western mediterranean sea). estuar. coast. 37:646-63. satta ct, pulina s, reñé a, padeddabm, caddeo t, fois n, lugliè a, 2020. ecological, morphological and molecular characterization of kryptoperidinium sp. (dinophyceae) from two mediterranean coastal shallow lagoons. harmful algae 97:101855. scholin ca, herzog m, sogin m, anderson dm, 1994. identification of groupand strain-specific genetic markers for globally distributed alexandrium (dinophyceae). 2. no nco mm er cia l u se on ly unruhdinium penardii in a mediterranean reservoir 83 sequence analysis of a fragment of the lsu rrna gene. j. phycol. 30:999-1011. smith vh, tilnan gd, nekola jc, 1999. eutrophication: impacts of excess nutrient inputs of freshwater, marine and terrestrial ecosystems. environ. pollut. 100:179-196. stamatakis a, 2014. raxml version 8: a tool for phylogenetic analysis and post-analysis of large phylogenies. bioinformatics 30:1312-1313. strickland jdh, parsons tr, 1972. a practical handbook of seawater analysis. fisheries research board of canada, ottawa, canada: 167 pp. takano y, hansen g, fujita d, horiguchi t, 2008. serial replacement of diatom endosymbionts in two freshwater dinoflagellates, peridiniopsis spp. (peridiniales, dinophyceae). phycologia 47:41-53. tardio m, tolotti m, novarina g, cantonati m, 2003. ecological and taxonomic observations on the flagellate algae characterizing four years of enclosure experiments in lake tovel (southern alps). hyrobiologia 502:285-296. utermöhl h, 1958. zurvervollkommung der quantitativenphytolankton-methodik. mitt. d. internat. vereinig. f. limnologie 9:1-39. watanabe mm, 2005. freshwater culture media, p. 13-21. in: r.a. andersen (ed.), algal culturing tecniques. elsevier accademic press. wu x, li c, chen l, zhao y, li y, wang h, 2018. responses of diel vertical migration behavior of peridiniopsis to the distribution of carbon and phosphorus in xiangxi river bay, three geoges reservoir. hu po ke kue 30:121-129. (in chinese with english abstract). xu y, cai q, wang l, kong l, li d, 2010. diel vertical migration of peridiniopsis niei, liu et al., a new species of dinoflagellates in an eutrophic bay of three-gorge reservoir, china. aquat. ecol. 44:387-395. you x, luo z, su y, gu l, gu h, 2015. peridiniopsis jiulongensis, a new freshwater dinoflagellate with a diatom endosymbiont from china. nova hedwigia 101:313-326 zhang h, lin s, 2005. phylogeny of dinoflagellates based on mitochondrial cytochrome b and nuclear small subunit rdna sequence comparisons. j. phycol. 41:411-420. zhang q, guoxiang l, hu z, 2011. morphological differences and molecular phylogeny of freshwater blooming species, peridiniopsis spp. (dinophyceae) from china. eur. j. protistol. 47:149-160. zhang x, gu h, luo z, sun j, 2013. description of peridiniopsis penardi (lemmerman) bourelly (peridiniales, dinophyceae) with its endosymbiotic diatom. trans. oceanol. limnol. 1:97104. (in chinese with english abstract). zhang q, liu g, hu z, 2014. description of a new freshwater bloom-forming dinoflagellate with a diatom endosymbiont, peridiniopsis minima sp. nov. (peridiniales, dinophyceae) from china. algol. stud. 145/146:119-133. no nco mm er cia l u se on ly layout 1 introduction microorganisms play a pivotal role in the marine carbon cycle (azam and malfatti, 2007), by producing, processing and utilizing one of the largest pools of active c-containing molecules on earth (hansell et al., 2009). in order to obtain the energy and carbon source required for growth, microbes utilize preferentially high-molecular weight substrates, as they are more bioavailable, according to the size-reactivity model (benner and amon, 2015). however, only small molecules can be taken up by the cells and therefore extracellular hydrolysis is a vital process for prokaryotic heterotrophs (reviewed by arnosti, 2011). in fact, measurable organic matter degradation rates have been found in all marine environments where life is possible, from deep subsurface sediments (hoarfrost et al., 2017) through the water column (hoppe and ullrich, 1999; celussi et al., 2018), to sea spray aerosol particles ejected to the atmosphere (malfatti et al., 2019), form the equator to the poles (misic et al., 2006; celussi et al., 2009). due to the wide array of environmental conditions (e.g. temperature, salinity, ph, pressure) where microorganisms thrive and produce active hydrolytic enzymes, such proteins harbor the potential for extensive utilization in biotechnological and industrial applications (nigam, 2013). specialized prokaryotes invest their energy efforts into the degradation of specific complex substrates (e.g. hydrocarbons, plastics; hassanshahian et al., 2014; yoshida et al., 2016) whereas generalists harbor the potential for hydrolyzing a wide array of macromolecules (martinez et al., 1996). another strategy is dependent on the location of exoenzymes with respect to the producing cell. in fact, exoenzymes can be cell-associated (i.e. attached to the cell wall or in the periplasmic space) or released (i.e. cellfree) in the surrounding waters (hoppe et al., 2002), thus identifying selfish vs ‘social’ foragers (traving et al., 2015). in the former case, after the uptake of large oligomers, further degradation occurs in the protected periplasmic space, avoiding diffusive loss of target monomers and enzymes (reintjes et al., 2017). in the latter case the release of enzymes is thought to be beneficial for microorganisms living attached to particles or colloids, where different cells can take advantage of specific degradation products made available by other consortium members in diffusionally constrained microenvironments (vetter et al., 1998). within this scheme reintjes et al. (2019) recently included the category of ‘scavengers’, defined as microbes that do not produce specific exoenzymes and take advantage of the ‘public goods’ made available by other cells. much attention has lately been addressed to cell-free enzymes, since they may not be strictly dependent on the metabolism of microbes, but article modulation of hydrolytic profiles of cell-bound and cell-free exoenzymes in antarctic marine bacterial isolates vincenzo manna,1,2 paola del negro,1 mauro celussi1* 1oceanography division, national institute of oceanography and applied geophysics (ogs), via auguste piccard 54, 34151 trieste; 2department of life sciences, university of trieste, via licio giorgieri 5, 34127 trieste, italy abstract microbial degradation of high molecular weight organic matter in the marine environment is dependent on the synthesis and activity of hydrolytic exoenzymes. these can be found both associated to the cell (i.e., attached to cell-wall or in the periplasmic space) and cell-free, dissolved in the surrounding waters. recent evidences suggest that exoenzymes location is linked to different foraging strategies. ‘selfish’ bacteria use cell-bound enzymes to obtain large oligomers which are then further degraded in the confined periplasmic space, whereas cell-free enzymes are thought to be advantageous for particle-attached prokaryotes, consequently ‘sharing’ the hydrolysis product with other members of the degrading consortium. we examined whether different bacterial isolates exhibit different exoenzymatic activity profiles when exposed to different growth condition. seven bacterial strains, isolated from the ross sea, were screened for the production of β-glucosidase, alkaline-phosphatase, lipase, chitinase and leucine aminopeptidase, assaying the hydrolysis rates of both cell-bound and cell-free exoenzymatic fractions. furthermore, to test whether bacteria growing on particles are more prone to produce cell-free exoenzymes, we set up a controlled experiment, amending one of the bacterial isolates with phytodetritus, in order to promote the colonization of aggregates. cell specific hydrolytic rates were highly variable and different isolates expressed different dominant hydrolytic activities. these results demonstrate a specialization for different substrates in different bacteria, further suggesting that some of the tested isolates may have an intrinsic potential to copiously produce cellfree exoenzymes. moreover, our results suggest that this specialization is not strictly genetically determined but varies according to growth conditions. the experiment performed with phytodetrital particles highlighted an increasing contribution of the dissolved exoenzymatic activity in samples bearing aggregates and associated bacteria. also, the tested isolate further modulated its hydrolytic machinery, expressing a different enzymatic profile when exposed to phytodetrital particles. the continuation of these experimental activities, testing the response of different bacterial isolates to different substrates, will help to determine the conditions under which different hydrolysis patterns develop, deepening the current knowledge on the organic matter cycling in the ocean. no nco mm er cia l u se on ly v. manna et al.34 their presence could also be a consequence of the history and dynamics of the water masses, due to their potential long-term persistence in the environment (baltar et al., 2013). within this framework we aimed at characterizing the degradative potential of several bacterial isolates and, by utilizing one of these model organisms, providing a proof of concept on the differential production of cell-free enzymes by particle-attached microbes, if compared to their free-living counterparts. we chose a set of bacteria isolated from several locations and depth in the ross sea (antarctica) for two main reasons. the first, of ecological nature, is related to the high summer productivity of the ross sea. indeed, in this system the export of particulate organic carbon (poc) to the mesopelagic is approximately 40-50% of the surface primary production (catalano et al., 2010), compared to a global estimate ranging from 1 to 40% (ducklow et al., 2001). since the fate of sinking poc is highly dependent on its mineralization by prokaryotes (>70% of the total loss during export; giering et al., 2014) we envisioned to obtain model organisms (bacterial isolates) and model particles (antarctic phytodetritus) to deepen the current knowledge on poc degradation and poc-bacteria interactions. the second reason is to be related to the biotechnological potential of bacteria isolated in extreme environments. due to the limited accessibility of these systems, the current knowledge on the bioactive molecules and secondary metabolites produced by extremophiles is extremely limited (lo giudice and fani, 2015), opening inestimable chances for novel discoveries. methods bacterial isolation the bacterial strains used for the experimental activities were isolated from water samples collected in austral summer 2006 and 2017. seawater was sampled at several locations and depths in the ross sea (tab. 1) by means of 12-l niskin bottles mounted on an sbe 32 carousel sampler. seawater aliquots were collected in sterile 50-ml tubes, plated onto zobell 2216 agar plates and kept at 4°c. bacterial colonies with different morphologies were selected and purified by serial streaking on the same medium for a minimum of four times. a total of 22 isolates (4 from 2006 and 18 from 2017 cruises) were maintained both in solid zobell 2216 agar plates and liquid zobell marine broth in the dark at 4 and 10°c, respectively. all the experimental activities described in this study were carried out between october and november 2018. bacterial identification from each isolate dna was extracted by means of the dneasy tissue kit (qiagen) according to the supplier’s tab. 1. details of sampling stations (coordinates, depth and location) in the ross sea and identifier of the strain isolated from each sample. isolates used for experimental activities are highlighted in bold. isolate id latitude (°s) longitude (°e) depth (m) date location af 74.9400 -176.2432 400 23/01/2006 glomar challenger basin, core of circumpolar deep water ag 74.9400 -176.2432 400 23/01/2006 glomar challenger basin, core of circumpolar deep water ak 74.9400 -176.2432 400 23/01/2006 glomar challenger basin, core of circumpolar deep water 9 74.7572 166.8105 910 23/01/2017 drygalki basin, bottom layer 9_1 74.7572 166.8105 910 23/01/2017 drygalki basin, bottom layer 9_2 74.7572 166.8105 910 23/01/2017 drygalki basin, bottom layer 14_1 74.7572 166.8105 30 23/01/2017 drygalki basin, deep chlorophyll maximum 14_2 74.7572 166.8105 30 23/01/2017 drygalki basin, deep chlorophyll maximum 14_3 74.7572 166.8105 30 23/01/2017 drygalki basin, deep chlorophyll maximum 15 74.7572 166.8105 5 23/01/2017 drygalki basin, surface layer 16_1 74.0000 175.089 572 17/01/2017 joides basin, bottom layer 16_1_1 74.0000 175.089 572 17/01/2017 joides basin, bottom layer 16_1_2 74.0000 175.089 572 17/01/2017 joides basin, bottom layer 16_1_3 74.0000 175.089 572 17/01/2017 joides basin, bottom layer 16_1_4 74.0000 175.089 572 17/01/2017 joides basin, bottom layer 16_1_5 74.0000 175.089 572 17/01/2017 joides basin, bottom layer 16_2 74.0000 175.089 572 17/01/2017 joides basin, bottom layer 20_1 74.0000 175.089 40 17/01/2017 joides basin, deep chlorophyll maximum 20_2 74.0000 175.089 40 17/01/2017 joides basin, deep chlorophyll maximum 20_3 74.0000 175.089 40 17/01/2017 joides basin, deep chlorophyll maximum 21 74.0000 175.089 5 17/01/2017 joides basin, surface layer no nco mm er cia l u se on ly cell-bound and cell-free enzymatic profiles of pelagic marine bacteria 35 instruction. the amplification of the 16s rrna gene was performed utilizing the universal primer 27f and the eubacterial primer 1492r as detailed in celussi et al. (2008). pcr products were purified using the qiaquick pcr purification kit (qiagen) according to the supplier’s instruction and amplicons were sequenced using abi prism big dye-terminator chemistry at the ‘bmr genomics’ facility at the university of padova (www.bmr-genomics.it). sequences were aligned to known sequences in the genbank database using blast (altschul et al., 1990). multiple sequence alignment was performed using the sina aligner (http://www.arb-silva.de/aligner/) and compared to the silva reference database release 111 (quast et al., 2013). phylogenetic tree was built using the phylogeny.fr (https://www.phylogeny.fr/; dereeper et al., 2008) tool, visualized and annotated with interactive tree of life (itol, v3; letunic and bork, 2016). the identical nucleotide sequence of several bacterial isolates led us to select 7 of them, out of the original 22, for the subsequent activities. the nucleotide sequences obtained in this study are deposited in genbank under the accession numbers mk780011-mk780031. exoenzymatic fingerprinting the selected isolates were transferred in a zobell marine broth diluted 5 times (zb/5) in 0.2 µm-filtered and autoclaved (121°c for 15 min) seawater (fasw) collected in the northern adriatic sea, and incubated at 0°c. after 2 days of acclimation at the experimental temperature, the isolates were inoculated in duplicate 100 ml aliquots of zb/5 at a starting abundance of 108 cell l–1 (d0). cell numbers were determined by flow cytometry according to the method described by marie et al. (1999). a facscanto ii (becton dickinson, franklin lakes, nj, usa) instrument was used, equipped with an air-cooled laser at 488 nm and standard filter setup. samples (0.5 ml) were ultrasonicated (10 cycles of 30 s on-off) and fixed with 0.5% (final concentration) glutaraldehyde (grade i for em analyses; sigma aldrich, st. louis, mo, usa). fixed samples were kept at 4°c for 15 min and diluted 1:10 with 0.2 μm-filtered tris-edta buffer 1x (sigma aldrich). samples were then stained with sybr green i nucleic acid dye (life technologies, carlsbad, ca, usa) at 1 x 10–4 dilution of the commercial stock and incubated for 10 minutes in the dark at room temperature. data were acquired using green fluorescence as trigger and processed with the facsdiva software (becton dickinson). the flow rate was calibrated daily, by running distilled water and weighing it before and after the run (at least 5 replicates). abundances were then calculated using the acquired cell counts and the respective flow rates. bacterial abundance was estimated every 2 days by means of flow cytometry, as described above. when the growth curves evidenced the exponential growth phase, 50ml of the cultures were washed twice in fasw by centrifuging at 3200×g for 10 min at 0°c. this procedure was aimed at transferring the isolates in a ‘poorer’ medium (1.48 mg l–1 of dissolved organic carbon; f. relitti, personal communication) and allowed the removal of hydrolytic enzymes in the dissolved phase. immediately after the washing with fasw, exoenzymatic activities (eeas) were measured (dexp). the activities of beta-glucosidase (bglu), alkaline phosphatase (ap), lipase (lip), chitinase (chit) and leucine aminopeptidase (ama) were estimated by the artificial fluorogenic substrate analogue method (hoppe, 1993). artificial substrates were added to 2 ml of the isolates in fasw in triplicate, at saturating concentrations. bglu, ap, lip, chit and ama were determined after the addition of 4-methylumbelliferone-β-d-glucoside (200 µm), 4-methylumbelliferone-phosphate (50 µm), 4-methylumbelliferone -oleate (100 µm), 4-methylumbelliferone n-acetyl-β-d-glucosaminide (200 µm) and leucine-7amino-methyl-coumarine (200 µm), respectively. all samples were incubated in the dark at 0°c for 3 h. enzyme activities were derived from the increase in fluorescence due to the cleavage of the artificial substrate, measured with a shimadzu rf1501 fluorometer at 365 nm excitation / 455 emission for bglu, ap, lip and chit and 380 nm excitation / 440 emission for ama. calibration curves were performed at each working session utilizing fasw and standard fluorochrome solutions (5 µm). the substrates and the standards were purchased from sigma-aldrich. the same procedure was performed on 0.2-µm-filtered subsamples to measure the activity of dissolved exoenzymes, that, as expected, was not detectable at this stage. the remaining volume of the cultures in fasw was incubated at 0°c for a number of days equal to the one necessary for bacteria to grow exponentially, as observed earlier during the experiment (i.e. from d0 to dexp, between 4 and 8 days, depending on the isolate). at the end of this incubation (dfin) eea measurements and cell counts were performed a second time. the hydrolysis rates (normalised per cell numbers) measured in the two time points (dexp and dfin) were then analysed with an nmds. the simultaneous analysis of several metabolic features (e.g. the degradation and/or utilization of specific substrates) provides useful information on the functional changes that microbes undergo in different culturing conditions and allows to visualize changes in the enzymatic profiles (celussi et al., 2008; del negro et al., 2018). the nmds was performed using a bray-curtis dissimilarity matrix, built with log-transformed (ln (x+1)) isolates’ hydrolysis rates. analysis of similarity (anosim, clarke and warwick, 1994) was used to test the significance of samples grouping on the nmds plot. nmds and anosim were carried out using the r package vegan (oksanen et al., 2018), under r 3.5.1 (r development core team, 2018). no nco mm er cia l u se on ly v. manna et al.36 exoenzymatic activities on phytodetrital particles in order to test whether bacteria growing on particles are more prone to produce cell-free (dissolved) exoenzymes, a second experiment was set up. the phytodetritus (particles) was generated from a microplankton net sample (20 µm mesh-size) collected in the ross sea (74.7572°s; 166.8105°e) on january 23rd, 2017 in the upper 100m of the water column. chlorophyll a concentration in the net sample was measured by the standard fluorometric procedure (lorenzen and jeffrey, 1980). the identification of microplanktonic organisms in the sample was carried out on a fixed aliquot (formaldehyde solution, 4% f.c.) at the inverted microscope (labovert fs leitz) equipped with phase contrast at a magnification of 400x and 630x, according to the utermöhl method (utermöhl, 1958). the net sample was subjected to 7 cycles of freezethawing (-80°c / +80°c, bidle and azam, 1999) after which cells were pelleted (20,000×g, 10 min, room temperature) and the supernatant (organic carbon in the dissolved phase) removed. cells were then resuspended in an equal volume of fasw. four 50-ml tubes containing fasw with 108 cell l–1 of the fast-growing isolate ag were amended with a volume of detritus to yield a concentration of 10 µg of chlorophyll a l–1, so as to simulate the particulate organic matter content during a phytoplankton bloom. the tubes were kept in a temperature-controlled circulating-water-bath at 10°c for 24 h to promote the colonization of particles by bacteria. after this step, the tubes were kept standing at the same temperature for additional 24 h to allow the settlement of particles at the bottom (martinez et al., 1996). afterwards, the upper 47 ml within the tubes, potentially bearing only freeliving bacteria and dissolved exoenzymes, were carefully transferred to a new sterile 50-ml tube and the remaining 3 ml (potentially bearing only particles and particle-attached bacteria) were added with fasw to a final volume of 50 ml. so, at the end of this process, 4 tubes contained particles+attached bacteria (part) and 4 tubes contained freeliving bacteria and the dissolved fraction of exoenzymes that have been produced in the previous 48 h (sur). half of the tubes were used for testing beta-glucosidase activity and the other half for testing leucine aminopeptidase activity, in experimental duplicates. eea were measured after 3 and 6 h (t3 and t6) from the splitting of particles and free-living bacteria, in order to allow the production of cell-free enzymes by particle-attached bacteria. eea were analysed as described above on the total and <0.2 µm fractions. at t3 and t6 samples for the enumeration of bacterial cells were collected and fixed with 0.2-µm filtered dolomite-buffered formalin (2% f.c.). in this experiment bacteria were counted at the epifluorescence microscope after dapi staining (as detailed in celussi et al., 2008) in order to discriminate between free-living and particle-attached cells. in this latter case, the number of cells were multiplied by a factor of 2 to compensate for the non-visible cells attached in the lower side of particles. results identification phylogenetic analysis of 16s rrna sequences identified the 22 bacterial isolates as members of the classes flavobacteriia and γ-proteobacteria (fig. 1). the identical nucleotide sequence of several bacterial isolates led us to select 7 of them (ag, 20_1, 16_1_1, 21, 9_2, ak and af), out of the original 22, for the subsequent experimental activities. exoenzymatic fingerprinting of bacterial isolates all the 7 selected isolates were able to produce the assayed exoenzymes (i.e., bglu, ap, lip, chit and ama) both at dexp and dfin, with the cell-specific activities of each enzyme varying up to 1 order of magnitude among different isolates (fig. 2). the enzymatic fingerprints of different isolates were quite variable, both in terms of degradative profiles and of relative contribution of cell-free and cell-bound exoenzymes to the total cell-specific activity. on the nmds plot (fig. 3), bacterial isolates were separated into three groups (anosim r=0.95, p=0.001), according to their prevalent exoenzymatic activity. isolates 21 and 16_1_1 were identified as mostly glycolytic strains (fig. 3), both at dexp and dfin (see section ‘methods’). at dfin these isolates showed the highest glycolytic activity among the selected bacterial strains, coupled with an increased contribution of the dissolved exoenzymatic fraction to the total specific activity (up to 100%, fig. 2). four out of the seven tested strains were identified as harbouring a high proteolytic potential (20_1, af, ak, ag; fig. 3), showing the highest ama specific hydrolysis rates, between 237.34±1.5 and 427.29±16.29 amol cell–1 h–1 (fig. 2). the contribution of the dissolved exoenzymatic fraction to the total specific ama activity was low for all the four ‘proteolytic’ strains (between 0 and 6%, fig. 2). while at dfin these strains were all identified as highly proteolytic, ak and ag were characterized by high production of lipases and alkaline phosphatases at the beginning of the experiment (dexp, fig. 3). this pattern was partly retained also in the at dfin as ak and ag were the only strains, among the proteolytic ones, to show a relatively higher specific alkaline-phosphatase and lipase activities (fig. 2). all the strongly proteolytic isolates showed extremely low values of bglu and chit specific activities (<2 and <1 amol cell–1 h–1, respectively, fig. 2). exoenzymatic activities on phytodetrital particles in sur tubes used to test bglu activity (see section ‘methods’), free-living bacterial abundance was no nco mm er cia l u se on ly cell-bound and cell-free enzymatic profiles of pelagic marine bacteria 37 1.40×109±3.21×108 cells l–1 at t3, remaining quite constant after 3 h (t6, 1.24×109±2.09×108 cells l–1, fig. 4a). particle-attached cells abundance was 2 order of magnitude lower, showing rather constant values at both sampling points (5.58×107±4.60×107 and 6.07×107±2.42×107 cells l–1 at t3 and t6, respectively, fig. 4a). tubes bearing particles+attached bacteria (part) were characterized by a higher abundance of particle-attached cells at both sampling points, showing a decreasing trend over time (3.70×108±1.82×107 and 2.41×108±7.23×107 cells l–1). contrariwise, free-living bacteria increased over time in the part treatment, from 6.37×108±5.92×107 to 9.85×108±1.26×107 cells l–1 (fig. 4a). a similar pattern was noticeable in the tubes used for fig. 1. neighbour-joining tree based on 16s rrna gene phylogeny of the 22 bacterial isolates, with methanocaldococcus fervens as outgroup. genbank accession numbers are given in parenthesis. isolates further used for experimental activities are highlighted in bold. the numbers above nodes denote bootstrap values; the scale bar indicates 0.01 changes per nucleotide. no nco mm er cia l u se on ly v. manna et al.38 ama activity assays (fig. 4). free-living bacteria in sur tubes showed constant abundance between t3 and t6 (1.17×109±1.65×108 and 1.26×109±6.23×107 cells l–1, respectively, fig. 4b) as well as particle-associated cells (6.80×107±1.83×107 and 5.12×107±8.9×106 cells l–1, respectively, fig. 4b). in part tubes, free-living bacteria increased from 6.08x108±2.49x107 on t3 to 9.14×108±2.18×107 cells l–1 on t6 (fig. 4b), while particle-attached bacteria abundance declined between sampling times (4.15×108±1.31×108 and 2.31×108±4.14×107 cells l–1, respectively, fig. 4b). results of bglu and ama cell-specific hydrolysis rates are reported in tab. 2. overall, the cell-free contribution to the total exoenzymatic activity was higher in tubes bearing aggregates and associated bacteria (part) at both sampling times and for both the exoenzymes tested (tab. 2). in sur tubes, specific cell-bound glycolytic activity increased by four-fold between t3 and t6 (0.30±0.05 and 1.30±0.47 amol cell–1 h–1, respectively, table 2), while the activity of the dissolved fraction mildly increased over time (tab. 2). this pattern resulted in diminished relative contribution of the cell-free to the total exoenzymatic activity, which decreased from 60.83±1.77 to 35.32±3.81% between t3 and t6 (tab. 2). tubes bearing aggregates and associated bacteria (part) were characterized by a higher activity of the dissolved exoenzymatic fraction, which remained quite constant over time (1.11±0.20 and 1.01±0.13 amol cell–1 h– 1 at t3 and t6, respectively; tab. 2). cell-bound specific activity increased from 0.46±0.08 to 0.85±0.03 amol cell–1 h–1 during the 3 hours incubation, reducing the cell-free profig. 2. cell-specific enzyme activities at dfin of 5 exoenzyme obtained from 7 marine bacteria isolates. on the top row, ama values are plotted on the right y-axis. error bars represent the standard deviation of three analytical replicates. bglu, β-glucosidase; ap, alkaline phosphatase; lip, lipase; chit, chitinase; ama, leucine amino-peptidase. tab. 2. cell-specific hydrolysis rates of β-glucosidase (bglu) and leucine amino-peptidase (ama) in amol cell–1 h–1 assayed during the experiment with phytodetrital particles. results are given as the mean and standard deviation of two experimental replicates. t3 t6 cell-bound cell-free % cell-bound cell-free % bglu sur 0.30±0.05 0.46±0.02 60.83±1.77 1.30±0.47 0.70±0.09 35.32±3.81 part 0.46±0.08 1.11±0.20 77.59±7.50 0.85±0.03 1.01±0.13 54.71 ±2.42 ama sur 206.04±48.83 41.18±6.26 16.73±0.77 192.03±28.12 32.33±3.26 14.43±0.35 part 101.32±3.18 30.54±1.30 23.18±1.55 106.67±5.95 31.20±1.26 22.67±1.89 sur, tubes with free-living bacteria + dissolved fraction of exoenzymes produced in the previous 48 h; part, tubes with particles + attached bacteria; cell-bound, cell-specific hydrolysis rates of cell-attached exoenzymes (total minus cell-free activity); cell-free, cell specific hydrolysis rates of dissolved exoenzymes (i.e., passing through a 0.2 μm filter); %, relative contribution of cell-free specific hydrolysis rates over the total exoenzymatic activity. no nco mm er cia l u se on ly cell-bound and cell-free enzymatic profiles of pelagic marine bacteria 39 fig. 3. non-metric multidimensional scaling (nmds 2 dimensional) plot of the selected isolates enzymatic profiles at dexp and dfin. arrows highlight the path of the samples between the two experimental conditions. the stress value and the anosim statistics are presented in the lower left and upper right corners, respectively. bglu, β-glucosidase; ap, alkaline phosphatase; lip, lipase; chit, chitinase; ama, leucine amino-peptidase. fig. 4. bar plots reporting bacterial abundance in the two experimental conditions (sur and part) over time, for tubes used to test a) bglu and b) ama activities. error bars represent the standard deviation from the mean of two experimental replicates. no nco mm er cia l u se on ly v. manna et al.40 portion of glycolytic activity from 77.59±27.50% at t3 to 54.71±2.42% at t6 (tab. 2). protease specific hydrolysis rates were two order of magnitude greater than bglu activity both in sur and part tubes. in samples potentially bearing only free-living microbes, both cell-bound and cellfree exoenzymatic activity decreased over time (tab. 2), with a rather constant proportion of dissolved to total ama between t3 and t6 (16.73±0.77% and 14.43±0.35%, respectively; tab. 2). over the 3-hours incubation, cell-specific protease activity showed little variation both for cell-bound (101.32±3.18 and 106.67±5.95 amol cell–1 h–1 at t3 and t6, respectively) and cell-free enzymes (30.54±1.30 and 31.20±1.26 amol cell–1 h–1 at t3 and t6, respectively), resulting in a constant contribution of the dissolved fraction over time (tab. 2). discussion in the ocean, bacteria experience heterogeneous concentrations of organic matter, including hot spots of particulate organic matter (i.e., living and dead phytoplankton cells and marine snow) as well as zones depleted in organic substrates (stoker, 2012). therefore, given this extreme variability, the ectohydrolytic profile of a bacterium must reflect a set of strategies to optimize the retrieval of organic substrates from the patchy spatial and temporal distribution of the organic matter in the ocean (martinez et al., 1996). the enzymatic profile of a single bacterium can thus be viewed as the expression of a peculiar degradative strategy and of its genetic identity, in nature as well as in culture. the bacterial isolates used in this study were grown in the same culture media, yet, notwithstanding the identical growth condition, they expressed a diverse degradative fingerprint. it is worth to point out that during the first experiment, bacteria were transferred from a rich (zb/5) to a poor medium (fasw), and since the enzymatic profiles were tested immediately after the washing of the cultures (dexp), it is very likely that the cells expressed enzymatic fingerprints adapted to high organic carbon concentrations. protease activity varied over 1 order of magnitude (between 42.65±3.79 and 427. 29±16.29 amol cell–1 h–1; fig. 2), although specific activity of bacteria that used to grow in a rich medium (dexp) was lower than the one tested after several days of growth in a poor one (dfin) (122.00±132.33 and 219.11±168.95 amol cell–1 h–1, respectively). since protease activity is inhibited or unaffected by the presence of low molecular weight substrate (donachie et al., 2001), is likely that the high concentration of readily accessible peptides in the rich medium (zb/5) suppressed the cell-specific protease expression, while the complex nature of natural organic matter in the ‘poor’ medium (i.e., natural doc), enhanced the proteolytic activity. the shift of the degradative fingerprint of 3 out of the 7 tested isolates between the two time points (fig. 3), partly confirm this hypothesis. indeed, the nmds plot highlighted a consistent change in the degradation pattern of isolates ag and ak, which were identified as strongly proteolytic in the ‘poor’ medium (figs. 2 and 3) whereas in the rich medium their specific proteolytic activity was ten-fold lower (44.78±1.49 and 23.40±1.71 amol cell–1 h–1, respectively; supplementary fig. 1). the same pattern was evident for isolate 9_2 (fig. 3), which almost doubled its cell-specific ama activity in the ‘poor’ medium (27.61±0.40 and 58.42±0.42 amol cell–1 h–1 for rich and ‘poor’ media, respectively, supplementary fig. 1 and fig. 2). these results agree with those reported by baltar et al. (2017), which found a tight coupling between ama activity and the complexity of organic matter supplied to the microbial community. while highlighting a certain degree of phenotypic plasticity for a subset of the tested isolates, the pattern showed by the nmds plot (fig. 3) also suggests that some of the bacterial isolates (i.e., 20_1, af, 16_1_1 and 21) are genetically ‘constrained’ to express a specific set of exoenzymes. it has been hypothesized that members of peculiar environments, like the high-latitude pelagic microbial community, may have streamlined their hydrolytic machinery, specializing in the degradation of those substrate that they are more likely to encounter (arnosti, 2014). this could be the case of isolates 16_1_1 and 21, which expressed the same hydrolytic pattern growing in both culture media (fig. 3). noteworthy, these two isolates expressed the highest cell-specific glycolytic activity, 1 order of magnitude higher than the other bacterial isolates (fig. 2). looking at the two growth conditions, glycolytic activity was, on average, faster in the rich medium relative to the ‘poor’ one (22.43±20.74 and 10.24±15.30 amol cell–1 h–1, respectively) as a consequence of the different concentration of substrates in the two media. in fact, polysaccharide hydrolysis is generally reported as substrate-inducible (sinsabaugh and follstad shah, 2012) and the zobell medium is richer in organic carbon (an in this case in polysaccharides, derived by the yeast extract; sommer, 1996) than natural seawater. these two strains showed the highest contribution of dissolved extracellular enzymes to the total specific hydrolytic activity, among the tested isolates (up to 100%, fig. 3). the active release of cell-free extracellular enzymes has been linked, among several other factors, to the response of marine bacteria to starvation (albertson et al., 1990). however, bacteria growing in the ‘poor’ medium experienced natural doc concentration (1.48 mg l–1), thus cell starvation was unlikely to occur. more likely, the enhanced contribution of the dissolved exoenzymatic fraction to the total cell-specific activity might be linked to the lifestyle of the tested bacterial isolates. indeed, the phylogenetic analysis identified the selected bacteria as members of the classes flavobacteriia and γno nco mm er cia l u se on ly cell-bound and cell-free enzymatic profiles of pelagic marine bacteria 41 proteobacteria (fig. 1), which are often found associated with particles (delong et al., 1993; teeling et al., 2012; crespo et al., 2013). modelling, experimental and field studies suggest that the production of dissolved exoenzymes could be advantageous to particle-attached bacteria (vetter et al., 1998; vetter and deming, 1999; ziervogel and arnosti, 2008). all our isolates expressed cell-free hydrolytic activity at dfin (fig. 2), although the contribution to total cell-specific activity varied greatly (0-100%, fig. 2). this pattern may derive from the appropriate substrate stimulation exerted by the natural doc supplied with the ‘poor’ medium (alderkamp et al., 2007). another possible explanation is that the isolates expressing the highest cellfree enzymatic activity (i.e., ag, 16_1_1 and 21, fig. 2), have the genetic potential to produce dissolved extracellular enzymes even when not attached to detrital particles. there are indications that the fate of an extracellular enzyme (i.e., to be cell-bound or cell-free) could be genetically predetermined (nguyen et al., 2019); thus, isolates expressing high cell-free exoenzymatic activity without the presence of particles could be viewed as ‘climax’ colonizers, extremely streamlined to take advantage of organic matter hot spots. given the expression of a conspicuous amount of cellfree extracellular enzymes in these isolates, we tested the hypothesis that the dissolved exoenzymatic activity may be enhanced by the presence of phytodetrital particles. for this test, one fast-growing bacterium, ag, identified as halomonas meridiana (fig. 1), was chosen. halomonas meridiana is a halotolerant, mesophilic, gram-negative bacterium, first isolated from antarctic hypersaline lakes (james et al., 1990). this species shows some adaptation to exploit particles as organic matter hotspots, such as flagellar motility, and the ability to produce metalloproteases, required for the functioning of adhesion or detachment mechanisms (anithajothi et al., 2014). our results show that the contribution of the dissolved fraction to the total specific activity was always higher in tubes bearing particles and attached bacteria (part, tab. 2), confirming our initial hypothesis. the difference between sur and part tubes was already evident after 3 h from the tubes splitting for both bglu and ama (tab. 2), albeit reducing over time. the reduced contribution of cellfree exoenzymes to total activity at t6 may be the result of the degradation of already present cell-free enzymes (i.e., produced during the colonization ad settling phases, see section ‘methods’), as these are themselves subjected to proteolytic hydrolysis (ziervogel and arnosti, 2008). the phytodetritus used for the experimental amendment was mainly composed of diatoms species (~80% of total microplankton on the net sample; f. cerino, personal communication), an important source of polysaccharides in the marine environment (ziervogel and arnosti, 2008; alldredge et al., 1993). nevertheless, glycolytic activity was extremely low (<2 amol cell–1 h–1, tab. 2) during our experiment, with values very similar to those measured during the first experiment (fig. 2). recent studies highlight the existence of an alternative polysaccharide uptake mechanism in marine bacteria (reintjes et al., 2017, 2019), suggesting that the use of simple model substrates (such as the one utilized in fluorometric assays) to assess glycolytic rates may be partially ineffective, especially when dealing with algal-derived, complex polysaccharides. the tested isolate, ag, also showed similar pattern of ama activity in sur tubes and when cultured in the ‘poor’ media, remarkably modulating proteolytic activity when exposed to phytodetrital particles. indeed, cell-specific hydrolysis rates in part tubes were ~50% lower than the rates measured both in sur tubes and expressed in its enzymatic profiles, with an enhanced contribution of cell-free ama activity in particle-enriched treatments (up to 23%), suggesting a fine regulation of the degrative machinery according to the environmental conditions. conclusions the results here presented showed that even closelyrelated bacterial isolates express different exoenzymatic profiles, and that the enzymatic fingerprint of an isolate may vary consistently according to the quality and quantity of the organic matter supplied as substrate for growth. moreover, we added a piece of evidence on the mechanisms regulating cell-free exoenzymatic activity, demonstrating that bacteria growing associated with phytodetrital particles are more prone to actively release dissolved extracellular enzymes. by investigating the ‘mise en place’ of the microbial hydrolytic machinery under different environmental stimuli, this work deepens the current knowledge of the factors driving the organic matter hydrolysis in the ocean, the rate-limiting step of the marine carbon cycle. the continuation of these experimental activities, testing the response of different bacterial isolates to different substrates, will help to determine the conditions under which different hydrolysis patterns develop, deepening the current knowledge on the organic matter cycling in the ocean. acknowledgements this study was carried out as part of the italian program for research in antarctica (pnra) with funds from the project priamo (prokaryotes interactions with antarctic phytodetritus: a microto macroscale voyage from the surface to the deep ocean), pnra16_00103. the constructive comments of f. malfatti about the experimental designs and those of two anonymous reviewers are greatly acknowledged. no nco mm er cia l u se on ly v. manna et al.42 corresponding author: mcelussi@inogs.it key words: organic matter hydrolysis; extracellular enzymatic activity; cell-free enzymes; particle-associated bacteria; free-living bacteria. received: 17 april 2019. accepted: 18 june 2019. this work is licensed under a creative commons attribution noncommercial 4.0 license (cc by-nc 4.0). ©copyright: the author(s), 2019 licensee pagepress, italy advances in oceanography and limnology, 2019; 10:8240 doi: 10.4081/aiol.2019.8240 references albertson nh, nystrom t, kjelleberg s, 1990. exoprotease activity of two marine bacteria during starvation. appl. environ. microbiol. 56;218-223. alderkamp ac, van rijssel m, bolhuis h, 2007. characterization of marine bacteria and the activity of their enzyme systems involved in degradation of the algal storage glucan laminarin. fems microbiol. ecol. 59:108-117. alldredge al, passow u, logan be. 1993. the abundance and significance of a class of large, transparent organic particles in the ocean. deep-sea res. i 40:1131-1140. altschul sf, gish w, miller w, myers ew, lipman dj, 1990. basic local alignment search tool. j. mol. biol. 215:403-410. anithajothi r, nagarani n, umagowsalya g, duraikannu k, ramakritinan cm, 2014. screening, isolation and characterization of protease producing moderately halophilic microorganism halomonas meridiana associated with coral mucus. toxicol. environ. chem. 96:296-306. arnosti,c. 2011. microbial extracellular enzymes and the marine carbon cycle. ann. rev. mar. sci. 3:401-425. arnosti c, 2014. patterns of microbially driven carbon cycling in the ocean: links between extracellular enzymes and microbial communities. adv. oceanogr. 2014:1-12. azam f, malfatti f, 2007. microbial structuring of marine ecosystems. nat. rev. microbiol. 5:782-791. baltar f, arístegui j, gasol jm, yokokawa t, herndl gj, 2013. bacterial versus archaeal origin of extracellular enzymatic activity in the northeast atlantic deep waters. microb. ecol. 65:277-288. baltar f, morán xag, lønborg c, 2017. warming and organic matter sources impact the proportion of dissolved to total activities in marine extracellular enzymatic rates. biogeochemistry 133:307-316. benner r, amon rmw, 2015. the size-reactivity continuum of major bioelements in the ocean. ann. rev. mar. sci. 7:185-205. bidle kd, azam f, 1999. accelerated dissolution of diatom silica by marine bacterial assemblages. nature 397:508-512. catalano g, budillon g, la ferla r, povero p, ravaioli m, saggiomo v, accornero a, azzaro m, carrada gc, giglio f, langone l, mangoni o, misic c, modigh m, 2010. the ross sea, p. 303-318. in: l. liu, k.-k. atkinson, l. quinones and r. talaue-mcmanus (eds.), carbon and nutrient fluxes in continental margins: a global synthesis. springer, new york. celussi m, balestra c, fabbro c, crevatin e, cataletto b, fonda umani s, del negro p, 2008. organic-matter degradative potential of halomonas glaciei isolated from frazil ice in the ross sea (antarctica). fems microbiol. ecol. 65:504-512. celussi m, cataletto b, fonda umani s, del negro p, 2009. depth profiles of bacterioplankton assemblages and their activities in the ross sea. deep-sea res. i oceanogr. res. pap. 56:2193-2205. celussi m, quero gm, zoccarato l, franzo a, corinaldesi c, rastelli e, lo martire m, galand pe, ghiglione jf, chiggiato j, coluccelli a, russo a, pallavicini a, fonda umani s, del negro p, luna gm, 2018. planktonic prokaryote and protist communities in a submarine canyon system in the ligurian sea (nw mediterranean). prog. oceanogr. 168:210-221. clarke kr, warwick rm, 1994. similarity-based testing for community pattern: the two-way layout with no replication. mar. biol. 118:167-176. crespo bg, pommier t, fernández-gómez b, pedrós-alió c, 2013. taxonomic composition of the particle-attached and free-living bacterial assemblages in the northwest mediterranean sea analyzed by pyrosequencing of the 16s rrna. microbiologyopen 2:541-552. del negro p, celussi m, de vittor c, fonda umani s, 2018. rapid acclimation of microbes to changing substrate pools in epipelagic waters of an antarctic polynya during austral summer 2003. polar biol. 41:1. delong ef, franks dg, alldredge al, 1993. phylogenetic diversity of aggregate-attached vs. free-living marine bacterial assemblages. limnol. oceanogr. 38:924-934. dereeper a, guignon v, blanc g, audic s, buffet s, chevenet f, dufayard j-f, guindon s, lefort v, lescot m, claverie j-m, gascuel o, 2008. phylogeny.fr: robust phylogenetic analysis for the non-specialist. nucleic acids res. 36:465-9. donachie sp, christian jr, karl dm, 2001. nutrient regulation of bacterial production and ectoenzyme activities in the subtropical north pacific ocean. deep-sea res. ii 48:17191732. ducklow h, steinberg d, buesseler k, 2001. upper ocean carbon export and the biological pump. oceanography 14:50-58. giering slc, sanders r, lampitt rs, anderson tr, tamburini c, boutrif m, zubkov mv, marsay cm, henson sa, saw k, cook k, mayor dj, 2014. reconciliation of the carbon budget in the ocean’s twilight zone. nature 507:480-483. hansell d, carlson c, repeta d, schlitzer r, 2009. dissolved organic matter in the ocean: a controversy stimulates new insights. oceanography 22:202-211. hassanshahian m, zeynalipour ms, musa fh, 2014. isolation and characterization of crude oil degrading bacteria from the persian gulf (khorramshahr provenance). mar. pollut. bull. 82:39-44. hoarfrost a, snider r, arnosti c, 2017. improved measurement of extracellular enzymatic activities in subsurface sediments using competitive desorption treatment. front. earth sci. 5:13. doi: 10.3389/feart.2017.000135. hoppe hg, 1993. use of fluorogenic model substrates for extracellular enzyme activity (eea) measurement of bacteria, no nco mm er cia l u se on ly cell-bound and cell-free enzymatic profiles of pelagic marine bacteria 43 p. 423-431. in: p.f. kemp, b.f. sherr, e.b. sherr and j.j. cole (eds.), handbook of methods in aquatic microbial ecology. crc press, boca raton. hoppe hg, arnosti c, herndl g, 2002. ecological significance of bacterial enzymes in the marine environment, p. 73-107. in: r.g. burns and r.p. dick (eds.), enzymes in the environment. activity, ecology, and applications. marcel dekker, new york. hoppe hg, ullrich s, 1999. profiles of ectoenzymes in the indian ocean: phenomena of phosphatase activity in the mesopelagic zone. aquat. microb. ecol. 19:139-148. james sr, dobson sj, franzmann pd, mcmeekin ta, 1990. halomonas meridiana, a new species of extremely halotolerant bacteria isolated from antarctic saline lakes. syst. appl. microbiol. 13:270-278. letunic i, bork p, 2016. interactive tree of life (itol) v3: an online tool for the display and annotation of phylogenetic and other trees. nucleic acids res. 44:242-245. lo giudice a, fani r, 2015. cold-adapted bacteria from a coastal area of the ross sea (terra nova bay, antarctica): linking microbial ecology to biotechnology. hydrobiologia 761:417-441. lorenzen c, jeffrey s, 1980. determination of chlorophyll in seawater. unesco technical paper in marine science no. 35. unesco, paris: 20 pp. malfatti f, lee c, tinta t, pendergraft ma, celussi m, zhou y, sultana cm, rotter a, axson jl, collins db, santander mv, anides morales al, aluwihare li, riemer n, grassian vh, azam f, prather ka, 2019. detection of active microbial enzymes in nascent sea spray aerosol: implications for atmospheric chemistry and climate. environ. sci. technol. lett. 6:171-177. marie d, brussaard cpd, thyrhaug r, bratbak g, vaulot d, 1999. enumeration of marine viruses in culture and natural samples by flow cytometry. appl. environ. microbiol. 65;45-52. martinez j, smith dc, steward gf, azam f, 1996. variability in ectohydrolytic enzyme activities of pelagic marine bacteria and its significance for substrate processing in the sea. aquat. microb. ecol. 10:223-230. misic c, castellano m, fabiano m, ruggieri n, saggiomo v, povero p, 2006. ectoenzymatic activity in surface waters: a transect from the mediterranean sea across the indian ocean to australia. deep-sea res. i 53:1517-1532. nguyen tt, myrold dd, mueller rs, 2019. distributions of extracellular peptidases across prokaryotic genomes reflect phylogeny and habitat. front. microbiol. 10:413. nigam ps, 2013. microbial enzymes with special characteristics for biotechnological applications. biomolecules 3:597-611. oksanen j, 2017. vegan: ecological diversity. r package. version 2.4-4. quast c, pruesse e, yilmaz p, gerken j, schweer t, yarza p, peplies j, glöckner fo, 2013. the silva ribosomal rna gene database project: improved data processing and webbased tools. nucleic acids res. 41:d590-d596. r development core team, 2018. r: a language and environment for statistical computing vienna, austria. reintjes g, arnosti c, fuchs bm, amann r, 2017. an alternative polysaccharide uptake mechanism of marine bacteria. isme j. 11;1640-1650. reintjes g, arnosti c, fuchs b, amann r, 2019. selfish, sharing and scavenging bacteria in the atlantic ocean: a biogeographical study of bacterial substrate utilisation. isme j. 13:1119-1132. sinsabaugh rl, follstad shah jj. 2012. ecoenzymatic stoichiometry and ecological theory. annu. rev. ecol. evol. syst. 43:313-343. sommer r, 1996. yeast extracts: production, properties and components. proceedings 9th int. symp. on yeasts, sydney, australia. stocker r. 2012. marine microbes see a sea of gradients. science 338:628-633. teeling h, fuchs bm, becher d, klockow c, gardebrecht a, bennke cm, kassabgy m, huang s, mann aj, waldmann j, weber m, klindworth a, otto a, lange j, bernhardt j, reinsch c, hecker m, peplies j, bockelmann fd, callies u, gerdts g, wichels a, wiltshire kh, glöckner fo, schweder t, amann r, 2012. substrate-controlled succession of marine bacterioplankton populations induced by a phytoplankton bloom. science 336:608-611. traving sj, thygesen uh, riemann l, stedmon a, 2015. a model of extracellular enzymes in free-living microbes: which strategy pays off? appl. environ. microbiol. 81:7385-7393. utermöhl h, 1958. [zur vervollkommung der quantitative phytoplankton-methodik].[article in german]. mitt. internat. ver. limnol. 9:1-38. vetter ya, deming jw, 1999. growth rates of marine bacterial isolates on particulate organic substrates solubilized by freely released extracellular enzymes. microb. ecol. 37:86-94. vetter ya, deming jw, jumars pa, krieger-brockett bb, 1998. a predictive model of bacterial foraging by means of freely released extracellular enzymes. microb. ecol. 36:75-92. yoshida s, hiraga k, takehana t, taniguchi i, yamaji h, maeda y, toyohara k, miyamoto k, kimura y, oda k, 2016. a bacterium that degrades and assimilates poly(ethyleneterephthalate). science 351:1196-9. ziervogel k, arnosti c, 2008. polysaccharide hydrolysis in aggregates and free enzyme activity in aggregate-free seawater from the north-eastern gulf of mexico. environ. microbiol. 10:289-299. no nco mm er cia l u se on ly layout 1 introduction lake iseo is one of the deepest lakes south of the alps, one of the largest italian lacustrine basins, and an important resource for agriculture, industry (manufacturing and building centers are its peculiarities), fisheries, and drinking water, as well as for recreation and tourism. its inflow and outflow are the oglio river, which is one of the major tributaries of the po river, and it crosses the po plain for about 156 km. the river is formed by the confluence of two alpine streams, one from the presena glacier in the adamello group, and the second from the corno dei tre signori (in stelvio national park). the water level of the lake is controlled by a dam built in 1933 at the outflow of the oglio river, which regulates water release in response to irrigation demand, while maintaining a sufficient flow in the river to sustain the health of downstream ecosystems (rotiroti et al., 2019). the lake is characterized by a high ratio between watershed area and lake surface (28.46); this feature, along with the mean annual precipitation value, which is about 900 mm y–1, strongly influences the characteristics of lake iseo, particularly the theoretical renewal time that is about four years (ambrosetti and barbanti, 1999) (fig. 1). considering its geographical position and morphological features, this lake is strongly influenced by increasing anthropogenic pressure and changes in meteoclimatic conditions. these aspects induce both direct and indirect impacts on the lake ecosystem. in effect, this lake should be a naturally oligotrophic water body; however, over the past 40 years, the increase of nutrient loadings brought it to a meso-eutrophic condition (leoni et al., 2014a; salmaso et al., 2014a). for an important system like the one of lake iseo, which provides many fundamental ecosystem services, continuous monitoring and reliable data to understand and predict physical, chemical, and biological evolution are crucial for social-welfare and management purposes. for this reason, since the 1970s, different research institutions have carried time series of limnological and ecological studies on lake iseo. they have worked independently or in synergy to improve knowledge on the ecosystems and to assess their dynamics. since the 1990s, research group of freshwater ecology (department of earth and environmental sciences, university of milano-bicocca) has sampled the lake on a monthly basis, analyzing the physical-chemical parameters of the lacustrine water, as well as the biotic community and the pelagic food web. monitoring programs have generated a wealth of data; details of parameters, sampling sites, depth, frequencies, sampling, and analysis methods are listed in tab. 1. these data have justified the inclusion of lake iseo in the italian and european longterm ecological research network (lter) and in the global lake ecological observatory network (gleon). the present study aims to synthesize the results of long-term investigations about the limnological state of lake iseo, providing updated data. major achievements obtained during several decades of research help us to outline knowledge gaps and future perspectives. all that is discussed in the context of climatic changes that act coupled with pre-existing stressors like eutrophication; these factors can work in synergy and reveal their effects on the lake ecosystem. changes observed in lake iseo were compared with those reported in the other deep south alpine lakes (dsl) to highlight analogies and differences of long-term chemical, physical, and biological patterns. in particular, we focus on the following aspects: i) the trophic evolution of the lake; ii) the thermal dynamic related to climatic changes that influences nutrient availability; iii) evolution in the phytoplankton community, with a special focus on cyanobacteria; iv) zooplankton review long-term studies for evaluating the impacts of natural and anthropic stressors on limnological features and the ecosystem quality of lake iseo barbara leoni, morena spreafico, martina patelli, valentina soler, letizia garibaldi, veronica nava department of earth and environmental sciences, university of milano-bicocca, piazza della scienza 1, 20126 milan, italy abstract we review the state of the art of limnological studies in lake iseo and provide updated data concerning long-term investigations (from 1993 to 2018) carried out on chemical and physical parameters (e.g., oxygen, phosphorus, silicon). changes observed in lake iseo were compared with those reported in other deep south alpine lakes (dsls) to highlight analogies and differences of long-term chemical, physical, and biological patterns. until the 1960s, lake iseo and other dsls were oligotrophic. the increase of anthropogenic pressure and global warming has led to a progressive and unrecovered process of eutrophication. moreover, the decrease in frequency of full mixing episodes has induced a state of temporary meromixis. other changes have been identified over the last two decades, especially concerning the phytoplankton and zooplankton communities, and new emerging chemical pollutants were detected. given the important ecological and socioeconomic role of lake iseo, long-term investigations are of paramount importance to understand the response of the lake ecosystem to climatic and anthropogenic stressors. these two factors can also act coupled with new combined and synergic effects. no nco mm er cia l u se on ly b. leoni et al.82 and planktonic food webs; and v) a mention to emerging pollutants and microplastics. trophic evolution throughout the last century, the main environmental issue concerning lake iseo was eutrophication. lake iseo should naturally be an oligotrophic phosphorus limited water body; however, over the past 40 years, the increase of nutrient loadings brought it to a meso-eutrophic condition (leoni et al., 2014a; salmaso et al., 2014a). after the 1960s, anthropogenic influences induced an acceleration of eutrophication that completely changed the ecological characteristics of this lake as well as of most of other dsls (barbieri and mosello, 1992). this process had negative effects on physical-chemical and biological features, as well as on food-web functions, e.g., decreasing the transparency of lake water, high oxygen concentration in the superficial layers, a marked decrease of oxygen in the hypolimnetic layers, and changes in the biovolume and structure of planktonic communities. starting from the 1970s, the whole coast was engaged in a project of sewage treatment, with the building of two treatment plants located at the northern and southern ends of the lake (garibaldi et al., 1999). the long-term dynamics of soluble reactive phosphorus (srp) reflects complicated variations linked to the stratification patterns and to the release of p from the sediments under anoxic conditions. mean annual srp concentrations for the whole water column increased from around 10 μg p l–1 in 1967, to 20, 32, and 50 μg p l–1 in 1973, 1980 and 1990, respectively (garibaldi et al., 1998). after 2001, concentrations progressively increased, reaching the maximum value of 77 µg p l–1 in 2018. the increase of p can also be explained by considering faulty sewage systems (barone et al., 2019) and the impact of external loads, of which the contribution should be taken into account for the evaluation of water quality (nava et al., 2019). indeed, as reported by garibaldi et al. (1999), a total load of 93 t y–1 of phosphorus was estimated in fig. 1. location, bathymetry and principal morphometric and hydrological characteristics of lake iseo. the dot indicates the position of the sampling point. no nco mm er cia l u se on ly responses to local and global stressors in lake iseo 83 1995-1997 from the main tributaries, principally deriving from nonpoint sources. the phosphorus trend observed in lake iseo is in contrast to those measured in other dsls (lakes garda, lugano, como, maggiore), since the measures put in place in these lakes for the reduction of nutrient loads allowed a progressive decrease of external phosphorus load (rogora et al., 2018). however, due to a reduced tendency to the total mixing of the whole water column (see following paragraph), hypolimnion conditions of some of these lakes (e.g., maggiore) are experiencing a decrease in dissolved oxygen (do) concentration that is likely to result in a release of phosphorus from sediment, consequently leading to an increase of this nutrient in the bottom layers (rogora et al., 2018). as long as mixing events do not occur, this phosphorus would be segregated at the lake bottom and not be available for primary producers. thermal dynamics, mixing depth, nutrient availability, and climate change lake iseo should be classified as warm monomictic, as water temperature never drops below 4°c. however, it has to be considered a holo-oligomictic lake; indeed, owing to the great depth, complete circulation of the water column does not occur every year, but only after harsh and windy winters. this event merely concerns the superficial layers of the lacustrine water, and in the last few decades, the lake has sporadically experienced complete vertical overturn (leoni et al., 2014a). thanks to the growing availability of physical data, the significant warming and increased stability of the water column of lake iseo have been documented after 1990 (ambrosetti and barbanti, 1999). in lake iseo, the tab. 1. list of the parameters collected in lake iseo since 1993 and description of instruments and methodologies adopted. parameter depth instrument period references transparency secchi disk 1993-ongoing dissolved oxygen(mg l–1) discrete sampling* wtw oxi 320 and multi 3630 ids 1993-ongoing ph discrete sampling* radiometer phm 83 1993-ongoing temperature(°c) discrete sampling mercury-filled celsius 1993-ongoing *continuous (0-250 m) reversing thermometer 2018-ongoing td-diver electrical conductivity(μs cm–1) discrete sampling radiometer cdm 1993-ongoing *continuous (0-250 m) td-diver 2018-ongoing alkalinity(ppm caco3) discrete sampling* acidimetric titration 1993-ongoing gran’s method total phosphorus(μg l–1) discrete sampling* spectrophotometer 1993-ongoing valderrama (1981) molybdenum reaction total nitrogen(μg l–1) discrete sampling* spectrophotometer 1993-ongoing valderrama (1981) (272/220nm) soluble reactive phosphorus(μg l–1) discrete sampling* spectrophotometer 1993-ongoing apha (1985) molybdenum reaction ammonia-nitrogen(μg l–1) discrete sampling* spectrophotometer 1993-ongoing apha (1985) indophenol reaction dissolved silicon(μg l–1) discrete sampling* spectrophotometer 1993-ongoing apha (1985) molybdenum reaction anions:no3–, cl–, so42–(mg l–1) discrete sampling* ion-chromatography 1993-ongoing manual, thermo fisher (dionex ics 1100) scientific in.c (2012) cations: ca2+, mg2+, na+, k+ (mg l–1) discrete sampling* ion-chromatography 1993-ongoing manual, thermo fisher (dionex ics 1100) scientific inc. (2012) chlorophyll-a continuous (0-20 m) zullig’s bottle 1993-ongoing lorenzen (1967) for integrate samples phytoplankton sampling continuous (0-20 m) zullig’s bottle 1998-ongoing for integrate samples phytoplankton analysis subsamples zeiss plankton im inverted 1998-ongoing lund et al. (1958) microscope using utermöhl’s smayda (1978) technique zooplankton sampling continuous (0-20 m) plankton nets (50-200 µm) 1998-ongoing zooplankton analysis subsamples zeiss plankton microscope 1998-ongoing de bernardi and canale (1995) *discrete sampling depths are the following: 0, 10, 20, 30, 50, 75, 100, 150, 200, 245 m. no nco mm er cia l u se on ly b. leoni et al.84 rate of temperature increase in the water column and/or hypolimnetic waters in the period between the 1970s and 2005/2009 ranged between 0.01 and 0.03°c y–1 (salmaso and mosello, 2010). the increasing trend of water temperature is common in dsls, as synchronous temporal patterns of water temperature were evident across the lakes, particularly since 2006, after the complete turnovers of 2005-2006, which caused a sudden cooling of the whole water column (rogora et al., 2018). during the period between 1986 and 2015, the warming trend in the surface waters of the lake iseo and other dsls was further evidenced through satellite data (pareeth et al., 2017). during the last four decades, mixing episodes involving the deepest waters of lake iseo occurred at the beginning of the 1980s, in 1999, 2000, 2005, and 2006. in other years, mixing depths ranged from 30 to 200 m with no regular pattern. in the most recent years (2014-2018), mixing depths reached less than 30% of the water column and less than 50% of lake volume (fig. 2; rogora et al., 2018). thermal stratification that began in spring was usually well-established between june and august, with thermocline depth varying from 10 to 20 m. owing to climatic factors (mild winters) and the accumulation of salts in the hypolimnion, lake iseo may currently be defined as temporary meromictic. the accumulation of salts mainly depends on the precipitation of calcium carbonate, which is a direct effect of high productivity in water rich in calcium and bicarbonate, as is the case with lake iseo (garibaldi et al., 1999). numerical simulations were applied to lake iseo to predict the effects of climate-change scenarios on lake thermal evolution (2012-2050). results showed an overall average increase in lake-water temperature (0.012°c y–1) and reinforced schmidt thermal stability of the water column in the winter, which may further hinder the deep circulation process (pilotti et al., 2013;valerio et al., 2015). a comparable result was obtained in the study of fenocchi et al. (2018), in which the application of the general lake model (glm) to lake maggiore suggested that persistent lack of complete mixing, severe water warming, and extensive effects on water quality are to be expected for centuries to come for all dsls. during the second half of the 1980s, there was a marked decrease in hypolimnetic oxygen concentrations, and anoxia conditions were established during the 1990s (garibaldi et al., 1999, 1995). the full mixing of 2005 and 2006 partially recharged dissolved oxygen in the water column, reaching around 60% saturation (leoni et al., 2014a). however, the effects of these events did not persist, and deep-water do again started to decrease soon afterwards. a hypoxic/anoxic condition was again established from 2010 onwards, with do concentrations in the deepest layers (>150 m depth) steadily below 1.2 mg o2 l–1. on the basis of the available data, the period from 2010 to 2018 represents the longest recorded phase of anoxia that lake iseo experienced (rogora et al., 2018). the volume-weighted average do concentration on the whole water column measured during the spring turnover showed stable values between 4 and 6 mg o2 l–1 in lake iseo; a slight tendency towards decreasing do levels was fig. 2. mixed depth as % of the maximum lake depth at winter-spring turnover in the lake iseo in the period 1993-2018. no nco mm er cia l u se on ly responses to local and global stressors in lake iseo 85 observed during the last five years, declining below 4 mg o2 l–1. in particular, anoxic conditions affect a growing hypolimnion volume due to a reduction of the renewal time of the deepest waters. these conditions promote phosphorus release from the sediment and the accumulation of reduced chemical species, such as sulphide, and methane, with severe consequences for the entire lake ecosystem (posch et al., 2012; rogora et al., 2015). these changes could have a huge impact on water usability, particularly for fishing, drinking, and recreational purposes. in lake iseo, tp concentrations gradually increased in the hypolimnion (150-251 m) after 2005, from about 60 to 90-100 µg p l–1, and even more in the deep-water layer (below 100 m: from 100 µg p l–1 during the period 1993-2005 to 140-160 µg p l–1 during 2014-2018 (rogora et al., 2018). a similar pattern to phosphorus emerged for hypolimnetic (150-251 m) dissolved silica (dsi) in lake iseo (fig. 3). indeed, dsi concentration was quite constant from 1993 to 2004, with average values (±standard deviation) of around 1509±92 µg si l–1. these values are higher than the ones observed for the same period in the superficial layers (0-20 m), where average values were 652±346 µg si l–1. in 2005, dsi concentration abruptly decreased, reaching a hypolimnetic value of 1300 µg si l–1 as a result of the complete circulation event that mixed bottom and superficial waters. however, after this abrupt decrease, dsi concentration gradually increased, reaching a value of about 2300 µg si l–1 in the most recent years. this increase can be linked to the change of the thermal behavior of lake iseo from oligomictic to meromictic, which is leading to the segregation of dissolved silica in the deepest layers. that may have consequences on diatom populations, dominating and supporting the pelagic community (see following paragraph), and, in a broader context, on the trophic web and productivity of lake iseo. in fact, several studies documented that the si internal loading from sediments can contribute over 40% of total annual dsi to lakes, and may thus be an important flux in the overall si budget and for sustaining the growth of siliceous phytoplankton (wang et al., 2016). the resistance to mixing, which causes detrimental effects to the deeper layers, could have apparent beneficial effects on surface waters and induce the reoligotrophication of the epilimnetic layers of the lake (fig. 4). however, considering lake iseo, even if the internal load is segregated at the bottom of the lake, external loads are enough to maintain high levels of primary productivity (barone et al., 2019). on the other hand, future complete mixing with the subsequent replenishment of the superficial layers with nutrients and reduced substances, may have detrimental effects on the entire lake ecosystem, as what happened in the past in lake lugano (holzner et al., 2009). many studies and synoptic analyses performed on longterm data (20-27 years) demonstrated that both the local climate (e.g., air temperature) and some large-scale climate patterns (e.g., teleconnection indices) impacted springwater temperatures and the mixing pattern of large and deep lakes south of the alps (manca et al., 2015; salmaso et al., 2014). climate variability in winter, thermal structure, mixfig. 3. trend of annual concentrations of dissolved silicon (dsi) in the hypolimnetic layers (150-251 m) of lake iseo from 1993 to 2018. no nco mm er cia l u se on ly b. leoni et al.86 ing regime, and the vertical redistribution of oxygen and nutrients in dsls appeared to be controlled by two inversely correlated modes of atmospheric circulation relevant for the mediterranean area: the east atlantic (ea) pattern and the eastern mediterranean pattern (emp; salmaso et al., 2014). in particular, the negative value of winter ea (or positive of winter emp) led to cold winterair temperature in the lake iseo area that caused deeper mixing during vertical spring turnovers and an increase of epilimnetic-nutrient (e.g., phosphorus) replenishment. however, a positive value of winter ea (or negative of winter emp) leads to warm winter-air temperature, a shallow mixing depth during vertical spring turnovers, and a segregation of hypolimnetic layers (leoni et al., 2018). phytoplankton community and cyanobacteria plankton communities are at the foundation of aquatic-ecosystem functioning, and an in-depth comprehension of their dynamics is fundamental to manage aquatic resources, and to predict and face future environmental changes. long-term series of plankton analyses provide essential datasets for characterizing reliable patterns of seasonal and multiannual trends, and shifts of populations and communities in response to global or local impact (morabito et al., 2018). standardized phytoplankton studies in the euphotic layers (0-20 m depth) of lake iseo have existed since the 1990s. in the last thirty years, about 60 phytoplankton taxa belonging to seven taxonomic groups were identified. at least three main algal groups dominated in terms of iseo’s biovolume community throughout the last three decades: bacillariophyceae, conjugatophyceae, and cyanobacteria (garibaldi et al., 2003; leoni et al., 2014b). studies of vertical phytoplankton distribution showed that spatial heterogeneity in the vertical dimension is much more significant than that in the horizontal dimension. indeed, results on phytoplankton biovolume and chl-a indicated that there is no significant variability in spatial distribution among sampling sites located some kilometers from each other (leoni et al., 2014a). phytoplankton can optimize their depth in the water column to maximize their growth rate using temperature, light, and nutrient gradients (reynolds, 2006). in summer, water temperature and light decrease with depth, while nutrients usually increase. in lake iseo, chemical parameters and inorganic nutrient concentrations showed a typical pattern for monomictic lakes, with low values of conductivity, phosphorus, nitrogen, and silica compounds in the surface layer in summer (diehl, 2002; minella et al., 2016; salmaso et al., 2012). a combination of field observations and 3d hydrodynamic simulations were used to identify phytoplankton species, and estimate the various time scales of the dominant physical and biological processes in lake iseo during a stratified period, deriving phytoplankton patch categorization and growth interpretation that provides a general framework for the spatial distrifig. 4. trend of annual concentrations of total phosphorus (tp) in surface (0-20 m) and hypolimnetic layers (150-251 m) of lake iseo from 1993 to 2018. no nco mm er cia l u se on ly responses to local and global stressors in lake iseo 87 bution of phytoplankton concentration in lake iseo (marti et al., 2015). during high water stability, vertical-transport time scales were much shorter than horizontal transport and vertical dispersion, leading to a growth niche for the observed phytoplankton. the large bacillariophyceae were dominant mainly during late winter and early spring (aulacoseira spp., melosira varians, asterionella formosa), conjugatophyceae (mostly mougeotia spp.) developed mainly in spring and summer, and cyanobacteria (mostly p. rubescens) developed in summer and autumn, following a fairly regular temporal trend during the years. furthermore, over the last few years, there have been several episodes of blooms of cyanobacterium dolichospermum lemmermannii, in particular, in conditions of high nutrient concentrations and high water temperature (salmaso et al., 2015). however, the described sequence can be integrated by the development of other algal taxa such as dinophyceae and chlorophyceae (garibaldi et al., 2003; hernández-avilés et al., 2018; leoni et al., 2014b). the greater importance assumed by the different algal groups in each year may be due to changes in meteoclimatic conditions during the growing season and to the different nutrient amounts recycled from the deep to the surface layers. in particular, in lake iseo, spring nutrient enrichment of epilimnetic layers, promoted by great vertical mixing depths due to harsh and windy winters, did not affect the increase of total phytoplankton biomass and chla concentrations, probably because an additional input of phosphorus in the already enriched epilimnion had minor effects (leoni et al., 2018). however, some effects on phytoplankton composition have been observed as, after deep mixing events, a great development of algal taxa with a high level of p content occurred (leoni et al., 2014a). during the last two decades, the predominant species in the cyanobacterial taxa was planktothrix rubescens. however, since 2014, the presence of “cryptogenic” species of cyanobacteria, tychonema bourrellyi, able to produce consistent biomasses and harmful toxins (anatoxin-a and homoanatoxin-a), was detected. stratified conditions in the aquatic system seem to be a key factor for the settling of t. bourrellyi, which cannot proliferate in lakes with deep mixing events. in lake iseo, enhanced stability triggered by higher air temperature promoted the development of these allochthonous cyanobacteria, which has created their niche as a result of the more stable water column and reduced mixing depths during the winter (nava et al., 2017). this important element of change is a good example of the taxonomic and ecological lack of knowledge on lake microbiota, and of the need for a more detailed list of potentially toxic species and the distribution of cyanotoxins in freshwater bodies (salmaso et al., 2018b). interesting insights for the monitoring of cyanobacteria and phytoplankton blooms in lake iseo, among other dsls, have been provided by the use of satellite data (bresciani et al., 2018, 2011). the first account on cyanotoxin distribution was carried out by cerasino and salmaso (2012) and updated in 2016 (salmaso et al., 2018a). in lake iseo, the conspicuous seasonal development of p. rubescens during spring and early summer corresponds to the increase of microcystin concentrations. instead, the coexistence of planktothrix and tychonema could explain the seasonal increase in anatoxin-a (salmaso et al., 2018b). zooplankton and food webs in lake iseo, the zooplanktonic community has been regularly sampled since the 1990s until today with monthly or bimonthly frequency. the huge quantity of collected data allowed to conduct long-term studies concerning the response of zooplankton-population dynamics to climatic variability and environmental changes, helping to clarify the effect that a change in zooplankton phenology can have on the entire lake pelagic food web. furthermore, this component is an early warning bioindicator of hydrophobic pollutants, as the zooplankton response to pollution changes resulted particularly prompt. the majority of research about zooplankton-population dynamics in lake iseo focused on the effects of vertical mixing dynamics on epilimnetic phosphorus replenishment and on plankton phenology. in particular, leoni et al. (2014a) investigated the effects of mixingdepth changes, during the late-winter and early-spring period, on the clutch size and population development of the daphnia longispina group, the fundamental primary consumers in the pelagic food web. the study highlighted the effects that the great replenishment of phosphorus in epilimnetic water levels, due to complete vertical mixing events, had on the growth rate of daphnia spp. the phosphorus increase in the surface-water layers seemed to predict the reproduction and population density of daphnia species, probably due to an increase in food quality (chlorophyta and large diatoms). winter climatic variability is a key factor in controlling vertical mixing depth and consequent phosphorus replenishment, triggering a cascading effect on lake limnological characteristics and on zooplankton phenology. leoni et al. (2018) verified that winter large-scale circulation patterns (ea and emp) control a chain of linked causal factors affecting winter-air temperature, spring-water temperature, resulting water vertical-mixing depth, and the epilimnetic concentration of total phosphorus and plankton phenology, including both primary and secondary consumers. this study indicated that, in lake iseo, characterized by high phosphorus concentration, a harsh winter corresponded to less favorable conditions for plankton development. in correspondence of a particuno nco mm er cia l u se on ly b. leoni et al.88 larly harsh winter, primary and secondary consumers postponed the time in which the population starts to increase after winter diapause, and the time in which primary consumers reach maximum density. this study demonstrated a different answer to climatic variability in eutrophic lakes such as lake iseo, compared to oligotrophic ones. indeed, in oligotrophic lakes, the increase of daphnia is triggered by an increase of phytoplankton growth deriving from spring enrichment in nutrients (manca et al., 2015). in lake iseo, spring enrichment in nutrients did not result in a considerable increase of phytoplankton growth, as phosphorus is always available in sufficient concentrations. thus, we hypothesized that the increase of zooplankton density can be linked to different phytoplankton quality instead of quantity (leoni et al., 2018; patelli et al., in press). the cascading effect triggered by winter-air temperature on springand summer-zooplankton phenology in lake iseo was also confirmed by patelli et al. (in press). the aim of this research was to disentangle the factors causing a daphnia population decline in summer (daphnia midsummer decline) in the lakes of lugano and iseo. winter-air temperature (through phosphorus replenishment) influenced daphnia abundance in spring in both lakes, but the effects were only propagated to summer daphnia abundance in lake lugano. additionally, summer daphnia abundance was influenced by summer-air temperature through a positive effect. results revealed some differences between the two study lakes, suggesting that, in lake iseo, which was characterized by stable mesotrophic conditions since the beginning of the 1990s, the decline of daphnia during summer is probably due to a physiological population decline after the spring peak. indeed, the intensity of the daphnia midsummer decline in this lake has very low interannual variability. the impact of the long-term chain of effects triggered by winter ea on daphnia were also documented for the lakes of garda and maggiore (manca et al., 2015). the effects of epilimnetic nutrient replenishment triggered by winter temperatures and plankton-biomass mixing were detectable in the meromictic lake lugano. despite being associated with lower p concentrations, warmer-thanusual winters and summers led to increases in summer chlorophyll-a concentrations through complex direct and indirect (food-web) effects (lepori et al., 2018). in lake iseo, a study about the zooplanktonic foodweb structure was conducted in order to detect the predator-prey relation considering organism-body size (leoni, 2017). in particular, the relation between the main predator, leptodora kindtii, and the potential prey (daphnia spp. and eubosmina longicornis) was reported by analyzing the stable isotope of c and n. size-specificity was crucial for addressing space and time changes in trophic links between organisms composing the two hierarchical levels within the open-water zooplankton community. indeed, among prey encountered by a free-swimming leptodora, only those able to fit into the basket opening could be captured. as basket diameter increases with animal-body length, prey size selection depends on l. kindtii body length. in particular, in spring, small-sized e. longicornis can be more suitable prey for l. kindtii than daphnia. cyclopoid adults were at the top of the zooplankton food chain in lake iseo, and they could potentially be feeding on daphnia. they, however, likely fed in a different habitat (>20 m deep water). emerging pollutants and microplastics industrial and agricultural production has contributed to the input of several pollutants in lake iseo. studies have focused on legacy (e.g., ddt, pcb) and current-use pops (e.g., polycyclic aromatic hydrocarbons, pahs) (binelli and provini, 2004, 2003; mazzoni et al., 2019; pascariello et al., 2019). salmaso and mosello (2010), and recently guzzella et al. (2018) in detailed studies reported the presence and concentrations of several organic and inorganic pollutants in lake iseo, as well in many other southern alpine lakes. several recent studies demonstrated the presence of micropollutants in different aquatic species ( beone et al., 2011; bettinetti et al., 2012a); some of these originated from the melting of glaciers that released pollutants accumulated in the past (bettinetti et al., 2012a). even in the case of lake iseo, an input of ddts from secondary sources was detected in the 2000s; sediment analyses confirmed that this load caused a drastic increase of ddts from 1991 to 2008, with higher values than those measured during the 1970s, when ddt was largely used in agriculture (bettinetti et al., 2011). a similar situation was observed in another dsl (maggiore; bettinetti et al., 2012b). an issue of emerging concern is microplastics, of which the presence in freshwater ecosystems is increasingly reported. in 2018, during the 10th edition of goletta dei laghi, a first monitoring of microplastics presence in the lakes of iseo, garda, and maggiore was carried out. results showed that microplastics (<5 mm) were found in all samples from the surveyed lake surfaces. the highest concentration was reported in lake iseo, with a value of 57,000±36,000 particles km–2 in the sampling transect performed between lovere and pisogne (sighicelli et al., 2018). knowledge gaps and future perspectives in these last few decades, lake iseo has experienced important and continuous changes attributable to both no nco mm er cia l u se on ly responses to local and global stressors in lake iseo 89 well-known stressors and emerging or never-before-considered types of impact. many studies were already published, monitoring activities are always in progress, and several other research topics have to be addressed. its responses are, in some cases, comparable to those globally reported in other lakes, but different local effects can also be highlighted. indeed, the complex interactions of different drivers (both climatic and anthropogenic) can affect aquatic systems on both a global and a local scale, on which the intrinsic characteristics of individual lakes also play a role. the complexity of these interactions has only been studied to a limited extent, and much is yet to be known about their single and/or synergic effects, and about lacustrine ecosystem responses. long-term investigation is an important component of our activities, giving us the power to detect and attribute the causes of environmental change, and to provide insights into how lakes function and react. however, further studies are still required to understand the legacy of processes that occurred in the past that are acting in combination with new and emerging stressors. in other dsls, long-term lake-monitoring data are now supplemented by high-frequency data collections that provide almost real-time information on lake condition and function, which is particularly valuable for lake modelling. to implement this technology on lake iseo, it could be beneficial to increase the ability to model and forecast long-term trends in lake thermal structure and p dynamics, crucial for developing reliable recovery plans. it is compelling to assess the role of climate change since climatic variations are somehow intensifying the symptoms of pre-existing negative types impact in freshwaters (moss et al., 2011). one of the most important open issues that strictly involve lake management is eutrophication. major efforts should be made to study catchment processes, especially focusing on the watershed load of nutrients on which lake communities rely, i.e., nitrogen and phosphorus. indeed, even if investigations were done to estimate the load linked to the watershed (barone et al., 2019; garibaldi et al., 1999), information is still scarce, and more and more detailed studies should be performed to obtain more reliable and accurate estimations. given hypolimnion segregation that is no longer mixed with lake surface water, catchment changes and processes can play a more important role on lake features. p and n concentrations primarily regulate total algal growth, but the relative availability of si to n and p can influence the composition of the phytoplankton community. despite this, the study of this important element is often dismissed. global projections on future climate predict an overall decrease in river discharge in southern europe (van vliet et al., 2013); therefore, since silica is introduced to the lake from inflowing surface waters, there could be a possible reduction in dsi concentrations in the surface water of lake iseo, as already verified in other systems, e.g., in lake baikal (sorokovikova et al., 2019). moreover, we already observed silica segregation in the deepest layers of lake iseo (see paragraph “thermal dynamics, mixing depth, nutrient availability, and climate change”). if the meromixis condition in this system becomes permanent, silica replenishment would not occur. this, coupled with decreased input from the watershed, results in a depletion of this important compound in superficial layers. a new investigation is being developed to model and understand the biogeochemical cycle of this compound that has an important effect on the phytoplankton community and, therefore, on the whole pelagic community. changes that occurred in the phytoplankton communities (such as the increased importance in lake iseo, as well as in the lakes of garda, maggiore, como, of tychonema bourrellyi, a cryptogenic species that had never before been observed, potentially toxic cyanobacterium that is radically changing phytoplankton features and the toxicological fingerprint of the largest dsls), underline the lack in knowledge concerning the taxonomy and ecology of lake microbiota, and the potential consequences for the ecological functioning of ecosystems, and for human health. furthermore, that highlights the need for developing next-generation monitoring approaches to use technological tools, such as the high-throughput sequencing (hts) of environmental dna. these innovative culture-independent approaches on the study of the biodiversity of prokaryotic and eukaryotic microorganisms in freshwaters, could radically modify our knowledge about the diversity, function, and biogeography of aquatic microbiota (salmaso et al., 2018a). remote sensing could help in monitoring the high spatial and temporal variability of some water-quality variables, such as those used in the evaluation of the ecological status of lakes according to the wfd. it provides frequent data for large-scale studies of water-quality parameters such as chlorophyll-a, which could significantly vary over short time periods and in the different areas of a lake (bresciani et al., 2018). zooplanktonic organisms, because of their relatively short life cycle, rapidly respond to local and global stressors. studies on the long-term data on lake iseo were crucial to understand changes taking place in the lake ecosystem; in particular, it was demonstrated that changes in phenology, such as growth rate, abundance, and species composition, can reflect climate shifts well. such changes can create a mismatch between consumers and food resources affecting the entire pelagic food webs. from this perspective, recent studies of food-web structure using stable isotopes highlighted the trophic relationship between taxa and the fate of hydrophobic pollutants in biota. the presence of micropollutants represents hidden and no nco mm er cia l u se on ly b. leoni et al.90 underestimated problems for aquatic biota and human health, and it is less perceived by citizens. in particular, the effects of emerging micropollutants (i.e., pfas) and mixtures of micropollutants are unknown, and they have an unpredictable impact level considering the thousands of toxic species released in hydrographic basins, lakes, and groundwater (cleuvers, 2003; ebele et al., 2017; valsecchi et al., 2015). the presence of microplastics in the surface water of lake iseo was verified (sighicelli et al., 2018). however, many studies highlighted that deep sediments can represent a long-term sink for these pollutants, so future investigations in lake iseo are needed to understand its role as a possible microplastics sink, through accumulation and segregation in deep sediments, or as a source, due to possible preferential distribution in the water column. a study is in place to disentangle the fate of microplastics in lakes, and to design a standardized method for the separation and analysis of microplastics from sediments, comparing the effectiveness of different methods and proposing a new one (nava et al., 2019). an important catchment-related aspect that it is often dismissed is the relationship between groundwater (gw) and surface waters (sw), since these systems create a hydrological continuum that could explain processes that would otherwise be difficult to understand. evaluating pathways and quantifying the fluxes between gw and sw systems are important for evaluating water-resource allocations. furthermore, a comprehensive understanding of gw-sw interactions supports the identification of pollutant migration pathways and, therefore, of their potential impact within the aquatic environment. as of now, this aspect has been marginally addressed, with a few studies performed at the basin scale (rotiroti et al., 2019). however, many aspects still remain to be considered, and studies could be performed to deeply understand the studied system due to its important implications on lake management. future studies on the lake iseo watershed could combine hydrodynamic, hydrochemical, and isotopic analyses of surface water and groundwater as an appropriate method for investigating the complex relationship between these compartments, and for evaluating anthropogenic hydrosystem modifications. in this contribution, we highlighted some important and continuous changes experienced by lake iseo, attributable to both well-known stressors and new emerging types of impact. we performed synoptic analysis providing updated data about the state of the art of long-term investigations and the most advanced studies carried out on this lake. we documented the effects of anthropogenic climate change on the lake ecosystem at different levels and scales, focusing on changes in physical and chemical features and on the distribution, abundance, and physiology of organisms, species interactions, and food webs. we highlighted ongoing studies and future perspectives to fill some knowledge gaps regarding lake ecological processes. to better understand key lacustrine dynamics and responses is essential for efficient management strategies to counteract climate change and mitigate its effects, and to conserve lake ecosystems, biodiversity, and natural capital. acknowledgments we would like to thank the collaborators and students who have contributed to the collection of historical data series. we are also very grateful to polizia provinciale di brescia and arpa lombardia-brescia for the help on sampling activities. investigations were partly carried out in the framework of the lter (long term ecological research) italian network, site southern alpine lakes, it08000-a (http://www.lteritalia.it/). researches on lake iseo have been funded by university of milano-bicocca (fa grant) and supported by national and regional grants. corresponding author: barbara.leoni@unimib.it key words: deep lake; eutrophication; plankton and food web changes; water column stability; pollutants. received: 17 october 2019. accepted: 13 december 2019. this work is licensed under a creative commons attribution noncommercial 4.0 license (cc by-nc 4.0). ©copyright: the author(s), 2019 licensee pagepress, italy advances in oceanography and limnology, 2019; 10:8622 doi: 10.4081/aiol.2019.8622 references ambrosetti w, barbanti l, 1999. deep water warming in lakes: an indicator of climatic change. j. limnol. 58:1-9. doi: 10.4081/jlimnol.1999.1 apha, 1985. standard methods for the examination of water and wastewater, 16th edition. apha, washington dc, us. barbieri a, mosello r, 1992. chemistry and trophic evolution of lake lugano in relation to nutrient budget. aquat. sci. 54:219-237. doi: 10.1007/bf00878138 barone l, milanesi l, valerio g, chapra sc, pilotti m, balistrocchi m, nizzoli d, 2019. analysis of the residual nutrient load from a combined sewer system in a watershed of a deep italian lake. j. hydrol. 571:202-213. doi: 10.1016/j.jhydrol. 2019.01.031 beone gm, cattani i, fontanella mc, ravera o, 2011. relationship between element concentrations and body size in the lake maggiore population of unio pictorum mancus (mollusca, bivalvia). j. limnol. 70:283-292. doi: 10.4081/jlimnol.2011.283 bettinetti r, galassi s, guilizzoni p, quadroni s, 2011. sedino nco mm er cia l u se on ly responses to local and global stressors in lake iseo 91 ment analysis to support the recent glacial origin of ddt pollution in lake iseo (northern italy). chemosphere 85:163-169. doi:10.1016/j.chemosphere.2011.06.037 bettinetti r, garibaldi l, leoni b, quadroni s, galassi s, 2012a. zooplankton as an early warning system of persistent organic pollutants contamination in a deep lake (lake iseo, northern italy). j. limnol. 71:335-338. doi: 10.4081/jlimnol.2012.e36 bettinetti r, quadroni s, manca m, piscia r, volta p, guzzella l, roscioli c, galassi s, 2012b. seasonal fluctuations of ddts and pcbs in zooplankton and fish of lake maggiore (northern italy). chemosphere 88:344-351. doi: 10.1016/j.chemosphere.2012.03.009 binelli a, provini a, 2003. the pcb pollution of lake iseo (n. italy) and the role of biomagnification in the pelagic food web. chemosphere 53:143-151. doi: 10.1016/s00456535(03)00441-7 binelli a, provini a, 2004. risk for human health of some pops due to fish from lake iseo. ecotoxicol. environ. saf. 58:139-145. doi:10.1016/j.ecoenv.2003.09.014 bresciani m, cazzaniga i, austoni m, sforzi t, buzzi f, morabito g, giardino c, 2018. mapping phytoplankton blooms in deep subalpine lakes from sentinel-2a and landsat-8. hydrobiologia 824:197-214. doi:10.1007/s10750-0173462-2 bresciani m, stroppiana d, odermatt d, morabito g, giardino c, 2011. assessing remotely sensed chlorophyll-a for the implementation of the water framework directive in european perialpine lakes. sci. total environ. 409:3083-3091. doi:10.1016/j.scitotenv.2011.05.001 cerasino l, salmaso n, 2012. diversity and distribution of cyanobacterial toxins in the italian subalpine lacustrine district. oceanol. hydrobiol. stud. 41:54-63. doi:10.2478/ s13545-012-0028-9 cleuvers m, 2003. aquatic ecotoxicity of pharmaceuticals including the assessment of combination effects. toxicol. lett. 142:185-194. doi: 10.1016/s0378-4274(03)00068-7 de bernardi r, canale c, 1995. ricerche pluriennali (19481992) sull’ecologia dello zooplancton del lago maggiore. doc. ist. ital. di idrobiol. 55:1-66. diehl s, 2002. phytoplankton, light, and nutrients in a gradient of mixing depths: theory. ecology 83:386-398. doi:10.1890/0012-9658(2002)083[0386:plania]2.0.co;2 ebele aj, abou-elwafa abdallah m, harrad s, 2017. pharmaceuticals and personal care products (ppcps) in the freshwater aquatic environment. emerg. contam. 3:1-16. doi:10.1016/j.emcon.2016.12.004 fenocchi a, rogora m, sibilla s, ciampittiello m, dresti c, 2018. forecasting the evolution in the mixing regime of a deep subalpine lake under climate change scenarios through numerical modelling (lake maggiore, northern italy/southern switzerland). clim. dyn. 51:3521-3536. doi:10.1007/ s00382-018-4094-6 garibaldi l, anzani a, marieni a, leoni b, mosello r, 2003. studies on the phytoplankton of the deep subalpine lake iseo. j. limnol. 62:177-189. doi:10.4081/jlimnol.2003.177 garibaldi l, brizzio mc, mezzanotte v, varallo a, mosello r, 1995. the continuing evolution of lago iseo (n.italy): the appearance of anoxia. mem. ist. ital. di idrobiol. 53:191212. garibaldi l, mezzanotte v, brizzio mc, rogora m, mosello r, 1999. the trophic evolution of lake iseo as related to its holomixis. j. limnol. 58:10-19. doi:10.4081/jlimnol. 1999.10 garibaldi l, mezzanotte v, brizzio mc, varallo a, mosello r, 1998. [apporti di fosforo al sebino. confronto fra misure sperimentali e teoriche].[article in italian]. acqua aria 98:105-110. guzzella lm, novati s, casatta n, roscioli c, valsecchi l, binelli a, parolini m, solcà n, bettinetti r, manca m, mazzoni m, piscia r, volta p, marchetto a, lami a, marziali l, 2018. spatial and temporal trends of target organic and inorganic micropollutants in lake maggiore and lake lugano (italian-swiss water bodies): contamination in sediments and biota. hydrobiologia 824:271-290. doi:10.1007/ s10750-017-3494-7 hernández-avilés js, callieri c, bertoni r, morabito g, leoni b, lepori f, buzzi f, salmaso n, 2018. prokaryoplankton and phytoplankton community compositions in five large deep perialpine lakes. hydrobiologia 824:71-92. doi: 10.1007/s10750-018-3586-z holzner cp, aeschbach-hertig w, simona m, veronesi m, imboden dm, kipfer r, 2009. exceptional mixing events in meromictic lake lugano (switzerland/italy), studied using environmental tracers. limnol. oceanogr. 54:1113-1124. doi:10.4319/lo.2009.54.4.1113 leoni b, 2017. zooplankton predators and preys: body size and stable isotope to investigate the pelagic food web in a deep lake (lake iseo, northern italy). j. limnol. 76:85-93. doi:10.4081/jlimnol.2016.1490 leoni b, garibaldi l, gulati rd, 2014a. how does interannual trophic variability caused by vertical water mixing affect reproduction and population density of the daphnia longispina group in lake iseo, a deep stratified lake in italy? inl. waters 4:193-203. doi:10.5268/iw-4.2.663 leoni b, marti cl, imberger j, garibaldi l, 2014b. summer spatial variations in phytoplankton composition and biomass in surface waters of a warm-temperate, deep, oligo-holomictic lake: lake iseo, italy. inl. waters 4:303-310. doi: 10.5268/iw-4.3.569 leoni b, nava v, patelli m, 2018. relationships among climate variability, cladocera phenology and the pelagic food web in deep lakes in different trophic states. mar. freshw. res. 69:1534-1543. doi:10.1071/mf17243 lepori f, bartosiewicz m, simona m, veronesi m, 2018. effects of winter weather and mixing regime on the restoration of a deep perialpine lake (lake lugano, switzerland and italy). hydrobiologia 824:229-242. doi:10.1007/s10750-0183575-2 lorenzen cj, 1967. determination of chlorophyll and pheopigments: spectrophotometric equations. limnol. oceanogr. 12:343-346. doi:/10.4319/lo.1967.12.2.0343 lund jwg, kipling c, le cren ed, 1958. the inverted microscope method of estimating algal numbers and the statistical basis of estimations by counting. hydrobiologia 11:143-170. doi:10.1007/bf00007865 manca m, rogora m, salmaso n, 2015. inter-annual climate variability and zooplankton: applying teleconnection indices to two deep subalpine lakes in italy. j. limnol. 74:123132. doi:10.4081/jlimnol.2015.1014 no nco mm er cia l u se on ly b. leoni et al.92 marti cm, imberger j, garibaldi l, leoni b, 2015. using time scales to characterize phytoplankton assemblages in a deep subalpine lake during the thermal stratification period: lake iseo, italy. water resour. res. 52:1762-1780. doi:10.1002/ 2015wr017555 mazzoni m, buffo a, cappelli f, pascariello s, polesello s, valsecchi s, volta p, bettinetti r, 2019. perfluoroalkyl acids in fish of italian deep lakes: environmental and human risk assessment. sci. total environ. 653:351-358. doi:10.1016/j.scitotenv.2018.10.274 minella m, leoni b, salmaso n, savoye l, sommaruga r, vione d, 2016. long-term trends of chemical and modelled photochemical parameters in four alpine lakes. sci. total environ. 541:247-256. doi:10.1016/j.scitotenv.2015.08.149 morabito g, mazzocchi mg, salmaso n, zingone a, bergami c, flaim g, accoroni s, basset a, bastianini m, belmonte g, bernardi aubry f, bertani i, bresciani m, buzzi f, cabrini m, camatti e, caroppo c, cataletto b, castellano m, del negro p, de olazabal a, di capua i, elia ac, fornasaro d, giallain m, grilli f, leoni b, lipizer m, longobardi l, ludovisi a, lugliè a, manca m, margiotta f, mariani ma, marini m, marzocchi m, obertegger u, oggioni a, padedda bm, pansera m, piscia r, povero p, pulina s, romagnoli t, rosati i, rossetti g, rubino f, sarno d, satta ct, sechi n, stanca e, tirelli v, totti c, pugnetti a, 2018. plankton dynamics across the freshwater, transitional and marine research sites of the lter-italy network. patterns, fluctuations, drivers. sci. total environ. 627:373-387. doi:10.1016/j.scitotenv.2018.01.153 moss b, kosten s, meerhoff m, battarbee rw, jeppesen e, mazzeo n, havens k, lacerot g, liu z, de meester l, paerl h, scheffer m, 2011. allied attack: climate change and eutrophication. inl. waters 1:101-105. doi:10.5268/iw-1.2.359 nava v, patelli m, leoni b, 2019. an r package for estimating river compound load using different methods. ecol. model. softw. 117:100-108. doi:10.1016/j.envsoft.2019.03.012 nava v, patelli m, soler v, leoni b, 2017. interspecific relationship and ecological requirements of two potentially harmful cyanobacteria in a deep south-alpine lake (l. iseo, i). water 9:993-1006. doi:10.3390/w9120993 pareeth s, bresciani m, buzzi f, leoni b, lepori f, ludovisi a, morabito g, adrian r, neteler m, salmaso n, 2017. warming trends of perialpine lakes from homogenised time series of historical satellite and in-situ data. sci. total environ. 578:417-426. doi:10.1016/j.scitotenv.2016.10.199 pascariello s, mazzoni m, bettinetti r, manca m, patelli m, piscia r, valsecchi s, polesello s, 2019. organic contaminants in zooplankton of italian subalpine lakes: patterns of distribution and seasonal variations. water 11:1901. doi:10.3390/w11091901 patelli m, leoni b, lepori f, in press. causes of daphnia midsummer decline in two deep meromictic subalpine lakes. freshwater biol. pilotti m, valerio g, leoni b, 2013. data set for hydrodynamic lake model calibration: a deep prealpine case. water resour. res. 49:7159-7163. doi:10.1002/wrcr.20506 posch t, köster o, salcher mm, pernthaler j, 2012. harmful filamentous cyanobacteria favoured by reduced water turnover with lake warming. nat. clim. chang. 2:809-813. doi:/10.1038/nclimate1581 reynolds cs, 2006. the ecology of phytoplankton, ecology, biodiversity and conservation. cambridge university press, cambridge: 552 pp. rogora m, buzzi f, dresti c, leoni b, lepori f, mosello r, patelli m, salmaso n, 2018. climatic effects on vertical mixing and deep-water oxygenation in the deep subalpine lakes in italy. hydrobiologia 824:33-50. doi: 10.1007/s10750-018-3623-y rogora m, mosello r, kamburska l, salmaso n, cerasino l, leoni b, garibaldi l, soler v, lepori f, colombo l, buzzi f, 2015. recent trends in chloride and sodium concentrations in the deep subalpine lakes (northern italy). environ. sci. pollut. res. 22:19013-19026. doi:10.1007/s11356-015-5090-6 rotiroti m, bonomi t, sacchi e, mcarthur jm, stefania ga, zanotti c, taviani s, patelli m, nava v, soler v, fumagalli l, leoni b, 2019. the effects of irrigation on groundwater quality and quantity in a human-modified hydro-system: the oglio river basin, po plain, northern italy. sci. total environ. 672:342-356. doi:10.1016/j.scitotenv.2019.03.427 salmaso n, albanese d, capelli c, boscaini a, pindo m, donati c, 2018a. diversity and cyclical seasonal transitions in the bacterial community in a large and deep perialpine lake. microb. ecol. 76:125-143. doi:10.1007/s00248-017-1120-x salmaso n, boscaini a, capelli c, cerasino l, 2018b. ongoing ecological shifts in a large lake are driven by climate change and eutrophication: evidences from a three-decade study in lake garda. hydrobiologia 824:177-195 doi:10.1007/ s10750-017-3402-1 salmaso n, boscaini a, capelli c, cerasino l, milan m, putelli s, tolotti m, 2015. historical colonization patterns of dolichospermum lemmermannii (cyanobacteria) in a deep lake south of the alps. adv. oceanogr. limnol. 6:1-4. doi:10.4081/aiol.2015.5456 salmaso n, buzzi f, cerasino l, garibaldi l, leoni b, morabito g, rogora m, simona m, 2014. influence of atmospheric modes of variability on the limnological characteristics of large lakes south of the alps: a new emerging paradigm. hydrobiologia 731:31-48. doi:10.1007/s10750-013-1659-6 salmaso n, buzzi f, garibaldi l, morabito g, simona m, 2012. effects of nutrient availability and temperature on phytoplankton development: a case study from large lakes south of the alps. aquat. sci. 74:555-570. doi:10.1007/s00027012-0248-5 salmaso n, mosello r, 2010. limnological research in the deep southern subalpine lakes: synthesis, directions and perspectives. adv. oceanogr. limnol. 1:29-66. doi:10.1080/ 19475721003735773 sighicelli m, pietrelli l, lecce f, iannilli v, falconieri m, coscia l, di s, nuglio s, zampetti g, 2018. microplastic pollution in the surface waters of italian subalpine. environ. pollut. 236:645-651. doi:10.1016/j.envpol.2018.02.008 smayda tj, 1978. from phytoplankters to biomass. in: a. sournia (ed.), phytoplankton manual. unesco, paris. sorokovikova lm, tomberg iv, sinyukovich vn, molozhnikova ev, khodzher tv, 2019. low water level in the selenga river and reduction of silica input to lake baikal. inland waters 9:464-470. doi:10.1080/20442041.2019. 1580078 valderrama jc, 1981. the simultaneous analysis of total nitrogen and total phosphorus in natural waters. mar. chem. 10:109-122. doi:10.1016/0304-4203(81)90027-x no nco mm er cia l u se on ly responses to local and global stressors in lake iseo 93 valerio g, pilotti m, barontini s, leoni b, 2015. sensitivity of the multiannual thermal dynamics of a deep pre-alpine lake to climatic change. hydrol. process. 29:767-779. doi: 10.1002/hyp.10183 valsecchi s, rusconi m, mazzoni m, viviano g, pagnotta r, zaghi c, serrini g, polesello s, 2015. occurrence and sources of perfluoroalkyl acids in italian river basins. chemosphere 129:126-134. doi:10.1016/j.chemosphere. 2014.07.044 van vliet mth, franssen whp, yearsley jr, ludwig f, haddeland i, lettenmaier dp, kabat p, 2013. global river discharge and water temperature under climate change. glob. environ. chang. 23:450-464. doi:10.1016/j.gloenvcha. 2012.11.002 wang b, liu cq, maberly sc, wang f, hartmann j, 2016. coupling of carbon and silicon geochemical cycles in rivers and lakes. sci. rep. 6:1-6. doi:10.1038/srep35832 no nco mm er cia l u se on ly layout 1 introduction many european lakes have been affected by a number of anthropogenic pressures, such as eutrophication, water level changes, toxic pollution and introduction of alien species. mountain lakes, relatively far from heavy populated areas, receive contaminants by long-range transport of air pollutants (rogora et al., 2008; poma et al., 2017). in the case of the monticchio lakes, even if they are relatively far from human activities, and enclosed in mt. vulture caldera, strong human impact was due to the touristic development in the 1960s, when houses, restaurants and a cableway were built, at present in large part abandoned. in order to follow the temporal trend in the ecological quality of ecosystems, long-term monitoring using well defined protocols can be used, providing deep insight in the changes in ecosystem structure and function (morabito et al., 2018). however, apart some rare exception (minder 1938) systematic monitoring programmes are often started after ecosystem alteration was evident (vollenweider et al., 1974; keating and dodd, 1975). scattered morphological, physical and biological observations are available for a number of lakes since the 1870s, concerning in particular the large lakes around the alps (marchetto, 1998), and the large volcanic lakes in central italy (margaritora, 1992). since the 1880s, a number of studies also deals with the monticchio lakes, and in this paper we evaluate the possibility to use them to infer the ecological history of these lakes during the periods when they were affected by climate changes, by change of nutrient levels in water, mainly related to the tourist pressure. a list of the studies used for this historical reconstruction is presented in the supplementary materials (tab. s1). the older papers (tata, 1778; palmieri and scacchi, 1852; de giorgi, 1879; cavanna, 1882; marinelli, 1895; vinciguerra, 1895; de lorenzo, 1900; casoria, 1901) only reported episodic information, frequently taken with rudimentary instruments. later on, a number of studies are related to one or few specific aspects of the lake water or biota, such as water chemistry (cannicci, 1952; squicciarini, 1974; mongelli et al., 1975), phytoplankton (cannicci, 1952; musacchio, 1981-1982), zooplankton (ruffo and stoch, 2005; alfonso, 2008), or macrophytes (venanzoni et al., 2003; azzella et al., 2010). however, four detailed limnological studies, covering most biological communities, were carried out. the first, on july 6th-8th, 1905, concerned lake morphometry, water chemistry, phytoplankton, zooplankton, macrophytes and phytobenthos and was reported by forti and trotter (1908) and stegagno (1908). in 1991, a second study concerned lake morphometry, water chemistry, phytoplankton, zooplankton (marano and d’aprile, 1991). after these two short campaigns, a two year study was carried out in 2005-2007 by the institute of ecosystem studies of the national research council (cnr-ise) (ceccanti et al., 2007), together with the milano-bicocca university, aiming to test the possibility to use the macrophytes to improve lago grande water quality and go into detail concerning water physical and chemical properties and article the contrasting evolution of two volcanic lakes lying in the same caldera (monticchio, mt. vulture, italy) inferred from literature records renato spicciarelli,1* aldo marchetto2 1school of agricultural, forest, food and environmental sciences (safe), university of basilicata, viale dell’ateneo lucano 10, 85100 potenza; 2cnr water research institute, corso tonolli 50, 28922 verbania-pallanza, italy abstract lago piccolo and lago grande di monticchio lie in the collapsed caldera of the volcanic structure of mt. vulture (basilicata, italy). in over two centuries, a number of studies on their water and on their submerged and riparian vegetation, were carried out, demonstrating an interesting biodiversity. the entire lake area, which is impacted by strong tourist pressure, is part of the “monte vulture” special area of conservation (sac it9210210). the aim of this paper is to review the literature studies on these lakes, in order to identify the more suitable limnological parameters to infer the history of the trophic status of the two lakes. for this reason, we assess the current ecological status of the two lakes on the basis of physical, chemical and biological analyses deriving from two recent surveys carried out in 2005-2007 and in 2015, and compare these data with sparse, but relevant, historical records, in order to assess how human impacts affected both these lakes and to understand the differences in their present trophic status. because of its peculiar water chemistry, lago piccolo is resulted in good and stable ecological conditions. on the contrary, water transparency of lago grande came out very low in summer, while total phosphorus and nitrogen concentration are proved high, leading to the persistence of critical environmental conditions in this lake, with high algal biomass and durable algal blooms in late summer, dominated by cyanobacteria. finally, in absence of standard protocols and seasonal samplings, the macrophyte maximum growing depth should be considered the more reliable indicator of trophic status among those available for these specific lakes, being relatively independent from sampling methods and seasonal pattern. no nco mm er cia l u se on ly literature record of monticchio lakes 45 phytoplankton and zooplankton communities. a last limnological campaign was carried out by the basilicata university (spicciarelli and mirauda, 2015), together with the umbria and marche experimental zooprophylactic institute and umbria regional agency for environmental protection. methods study area mount vulture is a composite volcano located in the basilicata region and formed by the superimposition of a number of volcanic edifices and affected by tectonic activity (la volpe et al., 1984). its formation started 800-750 kyears ago (principe, 2006). on the top of the mountain, at 650 m asl, a large caldera includes a tuff ring (giannandrea et al., 2006) and two maars (stoppa and principe, 1997), hosting the two monticchio lakes (fig. 1), divided by a rock strip and connected by a narrow channel through which water flows from lago piccolo (lp) to lago grande (lg). lp is smaller (0.155 km2) and deeper (44 m) than lg (0.41 km2 and 40 m), but the shape of lp is closer to a cone, while lg presents a large shallow area, ca. 10 m deep (fig. 2), so that the volume of lp (4.26 106 m3) is larger than the volume of lg (3.40 106 m3) (spicciarelli and mirauda, 2015). for lg, morphometric data obtained in 2015 compare well with those collected by stegagno (1908). lg sediment is mostly annually laminated and rich in tephras (wulf et al., 2004, 2008, 2012; schettler and alberic, 2008), and represents one of the best sedimentary records for paleoclimatic studies in the central mediterranean (watts, 1985; allen et al., 1999; brauer et al., 2007). the emerging watershed of the two monticchio lakes comprises approximately the caldera, with a surface of ca. 4 km2, mainly forested. beside precipitation in the catchment, monticchio lakes, and in particular lp, are also fed by a heterogenous volcanic aquifer, with alternatively high and low permeability structures and several underground water divides (celico and summa, 2004). from the volcanic structure, carbon dioxide reaches the lakes, and the quantity of co2 released compares with the two largest volcanoes producing co2 (gambardella, fig. 1. winter image of mt. vulture caldera with the two monticchio lakes, during lago piccolo (on the right) partial overturn. no nco mm er cia l u se on ly r. spicciarelli and a. marchetto46 2006), namely mt. etna (sicily) and popocatépetl (mexico). for this reason, lp is considered at risk of limnic eruption (chiodini et al., 1997, 2000; cioni et al., 2006; caracausi et al., 2009, 2013, 2015). in fact, a limnic eruption happened in 1810, and was described by ferdinando tortorella (1840) and quoted by another further visitor of mt. vulture (gussone and tenore, 1843; malpica, 1847; palmieri and scacchi, 1852; paci, 1853). human activity in mt. vulture caldera dates back to the 4th-3rd century bce, but a permanent settlement started in the 9th century, with the foundation of a basilian monastery. later a powerful benedictine abbey took over. anthropogenic pressure strongly increased in the 19th century, when forest surface was reduced because of tree cutting for using wood for railroad construction, and agriculture development, using lake water for irrigation. human pressure increased again in the 1950s when tourism activities developed, with a strong impact on the fig. 2. bathymetric maps of lago grande (left) and lago piccolo (right): top, in 1905 (stegagno, 1908); bottom, in 2015. reproduced from spicciarelli and mirauda, uep.o. fep 2007-2013 regione basilicata, 2015, with permission. no nco mm er cia l u se on ly literature record of monticchio lakes 47 ecosystems. in 1963, an endemic moth (brahmaea europaea hartig) living in the vulture caldera and in the close grotticelle wood was described, mainly feeding on narrow-leaved ash (fraxinus angustifolia subsp. oxycarpa (willd.)) (spicciarelli, 2004, 2015, 2018a, 2018b). it is considered a miocene relict by some authors. for this reason, the nature reserve of “grotticelle” was set in 1971, today included in the special area of conservation (sac) “grotticelle di monticchio” (spicciarelli, 2013). at present, the intensity of the tourism activities is lower, and natural vegetation is recovering, but the filling of the lg outlet in 2011 caused the increase of lake level, damaging ash forest around the lake. results temperature and transparency both lakes show a clear stratification in summer, with the thermocline located between 5 and 10 m, while the temperature is homogenous in winter (fig. 3). however, in lp there is an increase in temperature (up to 2.5°c) below 20 m, due to the presence of underwater springs. in the western part of lg, during summer and autumn the temperature between 4 and 7 m is 8-12°c lower than in the rest of the lake (spicciarelli and mirauda, 2015), probably as an effect of underwater springs. fig. 3. profiles of water temperature in lago piccolo (left) and lago grande (right) in april, june, september, december 2005 september data for lp were not available. redrawn from ceccanti et al., cnr-ise 2007, with permission. fig. 4. profiles of water ph in lago piccolo (left) and lago grande (right) in april, june, september, december 2005. redrawn from ceccanti et al., cnr-ise 2007, with permission. no nco mm er cia l u se on ly r. spicciarelli and a. marchetto48 during the recent studies, lg appeared warm monomictic, with full overturn in winter, as the volcanic lakes in central italy (margaritora, 1992). the same thermic regime was found by mongelli (1964, 1975), squicciarini (1974), and nicolosi et al. (2010). however, lg may freeze in winter, behaving as a dimictic lake (schettler and albéric, 2008). on the contrary, in spite of the homoeothermic conditions reached in winter, lp is meromictic (see below). several authors reported secchi disk depth measured in the monticchio lakes (tab. 1). comparing data from the end of the 19th century and the last decades, water transparency decreased in lg and increased in lp. dissolved gases and water ph underwater sources feed both monticchio lakes in carbon dioxide. in particular, in lp co2 concentration as high as 23 to 41 mmol l–1 were measured below the 20 m depth. in lg, values of 17-21 mmol l–1 were measured below the 20 m depth, with a maximum of 31 mmol l–1 close to the lake bottom in september and october 2008 (caracausi et al., 2013). such high co2 concentration affects water ph, which drops below 6 in the deeper parts of both lakes (nicolosi, 2010). at lake surface, water ph varies seasonally between 7 and 9 in both lakes (fig. 4), depending on the intensity of algal photosynthesis (dumontet et al., 2003; ceccanti et al., 2007), which also affect oxygen content. in fact, during summer the high photosynthetic activity leads to o2 oversaturation (dumontet et al., 2003). in lp, o2 concentration increases in summer from 11-12 mg l–1 at the surface to 23-25 mg l–1 at 5 m depth, dropping to zero below 20 m (ceccanti et al., 2007; spicciarelli and mirauda, 2015). in lg, the seasonal pattern is similar, but the depth of water anoxia fluctuates between 5 m in summer and 20 m in winter (fig. 5), when a large portion of lake bottom, lying at 12 m depth, is oxygenated. a seasonal pattern also affects methane concentration in lg, but not in lp: nicolosi (2010) reports that in the lp, the lowest methane concentration was measured in shallow tab. 1. reported secchi disk depth (sd) and maximum macrophyte growing depth (mmgd) in lago grande and lago piccolo. date lago grande lago piccolo source sd (m) mmgd (m) sd (m) mmgd (m) july 1905 3.3 7 4.3 4 forti and trotter, 1908; stegagno, 1908 jun-sep 1991 1.0-1.5 3-5 marano and d’aprile, 1991 summer 1994 0.3 “high” schettler and alberic, 2008 may 2001 0.5-1.0 3-3.5 dumontet et al., 2003 february 2005 1.9 ceccanti et al., 2007 august 2010 1.4 3.3 6.2 6.2 azzella et al., 2010 september 2015 1.0-1.2 3.2-3.4 4.7-7 5-6 spicciarelli and mirauda, 2015 fig. 5. profiles of dissolved oxygen concentration in lago piccolo (left) and lago grande (right) in april, june, september, december 2005. redrawn from ceccanti et al., cnr-ise 2007, with permission. no nco mm er cia l u se on ly literature record of monticchio lakes 49 waters, increasing towards the lake bottom in the monimolimnion, without a marked seasonal pattern. in lg, methane increases with depth and the highest concentrations in the entire hypolimnion have been measured during summer and autumn, while lower values were found in winter and spring. water chemistry the most striking aspect of the ionic composition of lp water is the very high iron concentration (fig. 6), reaching values higher than 100 mg l–1 in the anoxic deep water and causing lake meromixis (ceccanti et al., 2007; spicciarelli and mirauda, 2015). high values of mn (up to 2 mg l–1 in lp), ba and sr were measured in anoxic water in both lakes (tab. 2). such high iron concentration, coupled with the different thermal regime of the two lakes, dramatically affects the seasonal pattern of phosphorus concentration (fig. 7). in lp, iron concentration in oxygenated water causes the formation of insoluble fe3+ salts, which settle towards the deepest part of the lake, so that total p concentration in the upper 20 m of the lake never reaches 0.02 mg l–1. in lp anoxic bottom water, fe3+ is reduced to fe2+, and p is released, reaching a concentration of 0.9 mg l–1 in june 2005 (ceccanti et al., 2007). in lg iron concentration is lower, and epilimnetic fig. 6. profiles of iron concentration in lago piccolo (left) and lago grande (right) in april, june, september, december 2005. note the different scale. redrawn from ceccanti et al., cnr-ise 2007, with permission. fig. 7. profiles of total p concentration in lago piccolo (left) and lago grande (right) in april, june, september, december 2005.note the different scale. redrawn from ceccanti et al., cnr-ise 2007, with permission. no nco mm er cia l u se on ly r. spicciarelli and a. marchetto50 total p concertation reaches 0.06 mg l–1. during summer, in the anoxic hypolimnion total p concentration increases to values over 2 mg l–1 (ceccanti et al., 2007; spicciarelli and mirauda, 2015). winter overturn causes on one hand a small increase in bottom o2 concentration, causing a drop in total p concentration. on the other hand, water mixing leads to a p flux from deep water to surface water. during summer, total p concentration increases, suggesting the presence of local sources, such as the touristic activities on the lake shore. total nitrogen profiles (fig. 8) compare with the total phosphorus ones. most inorganic n is in the form of nitrate in oxygenated water and of ammonium in anoxic water. epilimnetic total n concentrations (0.5 to 3 mg l–1) are high enough to assure that n is not a limiting factor for algal growth (teubner & dokulil, 2002). tab. 2. range of metal concentration in lago grande and lago piccolo in 2005. metal (µg/l) lago piccolo lago grande min max min max al 5.25 55.9 7.8 182 as <dl 6 <dl 6.8 b 42.7 483.3 24.7 74.2 ba 58.4 424.85 69 289 cd <dl 11.7 <dl 1 co 0.1 434 <dl 20 cr <dl 0.9 <dl 0.9 cu <dl 2.45 <dl 3.75 fe 20.55 131055 14.25 5900 li 4.5 7.3 1.4 3.6 mn 1.35 2031.5 72.9 3688.5 ni <dl 4.5 <dl 2.75 pb <dl 20.2 <dl 2.5 pt <dl 82.7 <dl 11.5 se <dl 3.2 <dl 8.7 sr 256.65 605.15 363.9 630 tl <dl 8.7 <dl 9.3 v 14.8 46 21.2 38.7 zn 0.25 15.2 <dl 17.7 <dl, lower than detection limit. fig. 8. profiles of total n concentration in lago piccolo (left) and lago grande (right) in april, june, september, december 2005. redrawn from ceccanti et al., cnr-ise 2007, with permission. no nco mm er cia l u se on ly literature record of monticchio lakes 51 macrophytes an increase in the maximum macrophyte growing depth (mmgd) reached by submerged vegetation in lp, paralleled by a decrease in mmgd in lg, were recorded, in good agreement with changes in water transparency discussed above. in fact, in 1880 biondi reported the presence of a number of macrophytes, such as ceratophyllum demersum l., potamogeton natans l., p. pectinatum l., p. crispus l., myriophyllum verticillatum l., sparganium ramosum huds., conium maculatum l. (cavanna, 1882). in 1905, trotter detailed the distribution of the vegetation, finding 3 vegetation belts in lg and two in lp (forti and trotter, 1908). however, in the recent studies performed by venanzoni et al. (2003) and azzella et al. (2010), macrophyte cover appeared simplified in both lakes, composed by wet meadows and reed beds of carex, juncus and phragmites, because of the strong human impact on the shore habitat. detailed analysis of macrophyte cover using aerial photographs showed an increase in the presence of nymphaea alba l. and a decrease in potamogeton lucens l. and myriophyllum spicatum l. furthermore, taxodium distichum (l.) rich. is invading lake shores, causing a reduction in the presence of fraxinus angustifolia subsp. oxycarpa, an important species forming the main food source for the endemic moth brahmaea europaea (spicciarelli and mirauda, 2015; spicciarelli et al., 2016). plankton a number of field campaign on lakes monticchio were performed from 1905 to 2015 (see tab. s1). however, results are not directly comparable, as forti and trotter (1908) and cannicci (1952) collected phytoplankton samples using nets, while later studies used sampling bottles and the utermöhl technique (1952). concerning the more recent studies, ceccanti et al. (2007) collected samples in lg monthly in 2005 and bi-monthly in 2006 and they were able to describe the seasonal pattern of phytoplankton assemblages. on the contrary, musacchio (19811982) and marano and d’aprile (1991) collected samples in both lakes, only in one date (in june 1981 and in october 1991, respectively), while spicciarelli and mirauda (2015) in two dates (july and october). in lp ceccanti et al. (2007) only collected one sample in march 2005. in lp, all studies agree in describing a typical oligotrophic assemblage, dominated by diatoms and dinophytes or chlorophytes, with low algal abundance. on the contrary lg phytoplankton dominant species belonged to cyanobacteria, while the higher biomass was formed by chloropythes. in 2005, cyanobacterial assemblages were dominated in spring by woronichinia naegeliana (unger) elenkin and dolichospermum planctonicum (brunnthaler) wacklin, l.hoffmann & komárek, and in summer and autumn by microcystis aeruginosa (kützing) kützing and romeria leopoliensis (raciborski) koczwara, while merismopedia trolleri bachmann was present in both periods (ceccanti et al., 2007). in 2006, a summer bloom of aphanocapsa delicatissima west & g.s.west with the presence of snowella atomus komárek & hindák and anagnostidinema amphibium (c.agardh ex gomont) strunecký, bohunická, j.r.johansen & j.komárek (ceccanti et al., 2007). in lg, the maximum abundance value was reached in august 2006 (ca. 430 106 cells l–1), while in the other studies lower abundances values were detected, from 150 to 230 106 cells l–1, but samples were not collected in august. the monthly distribution of the measured phytoplankton abundance is reported in fig. 9, showing the importance of the choice of the sampling period for evaluating phytoplankton biomass. results of marano and d’aprile (1991) apparently did not include small cyanobacteria, so that the total density resulted very low (20.5, 20.2 and 15.5 106 cells l–1 in spring, summer and autumn, respectively). in the case of zooplankton, sampling and analysis were more consistent among the studies. at the beginning of the 20th century, forti (1908) found both lakes dominated by rotifers. unfortunately, marano and d’aprile (1991) did not considered rotifers, but ceccanti et al. (2007) and spicciarelli and mirauda (2015) found again their numerical importance in the zooplankton assemblages, while biomass was dominated by cladocera. most zooplankton species were found in both lakes (tab. s2), but assemblages were different in the two lakes: lp hosted a typical lacustrine community, while in lg species typical of pond habitats were found (spicciarelli and mirauda, 2015). ceccanti et al. (2007) also found the zooplankton biomass increased in spring, falling down in july because of both cyanobacterial blooms and the predation by chaoborus flavicans (meigen) and c. crystallinus (de geer) (leoni and garibaldi, 2009). fauna after the collection and description of the endemic bleak alburnus albidus o. g. costa (tenore, 1844), the first systematic description of monticchio fauna was edited by cavanna (1882), involving a number of specialists, and reporting lists of arachnidae, myriopoda, exapoda, pisces, amphibia, reptilia, aves, mammalia, mainly related to the aquatic environment. on the riparian hygrophytes, they found molluscs, such as planorbis umbilicatus (müller), lymnaea peregra (o. f. müller), l. lagotis schrank, ancylus latus edwards, succinea megalonixia bourg, dragonflies, auch as anax formosus v. d. lind., agrion tenellum dev., a. elegans v. no nco mm er cia l u se on ly r. spicciarelli and a. marchetto52 d. lind., crocothemis erythraea brullè, and a large number of platycnemis pennis pall., a species living on floating plants. other interesting records include the bird ardeola ralloides scopoli, the amphibian bombinator igneus merr. and the reptilian tropidonotus natrix l. among fishes, in the monticchio lakes were found the endemic bleak, the carp (cyprinus carpio l.), the european perch (perca fluviatilis l.), the black bass (micropterus salmoides lacépède), the italian chub (squalius squalus l.), the tench (tinca tinca l.), the catfish (ameiurus melas rafinesque), rutilus aula bonaparte, the eel (anguilla anguilla l.), the crucian carp (carassius carassius l.) and the western mosquitofish (gambusia affinis baird and girard). in lp, spicciarelli and mirauda (2015) also found alien species, such as the brown trout (salmo trutta l. morpha fario), and some specimen of japanese koi and white carps. discussion to understand the response of lake ecosystems to anthropogenic pressure and global change, the best approach is the long-term (multi-decadal) collection of limnological data with uniform protocols of sampling and analysis, such as those obtained for lake maggiore (morabito et al., 2005), lake constance (häse et al., 1998) or lake geneva (tadonléké et al., 2009). however, for a relevant number of lakes, a series of short-term episodic studies were carried out in the past, frequently by different authors, using different methods. frequently, such studies predate the starting of regular series of data collection. for example, in italy studies on lake maggiore (baldi et al., 1953), garda (d’ancona et al., 1961) or trasimeno (moretti, 1958) were carried out far before the start of the long-term ecological studies currently running (morabito et al., 2018). lakes monticchio are an example of lakes for which a large number of studies exist, frequently unpublished or difficult to find. they were mostly carried out independently, with no effort to obtain comparable data among them. in this paper, we try to assess if this kind of studies can be used to infer the trophic history of these lakes. from the different studies, it emerges that the main effect of increase in human activities starting in the 1960’s was eutrophication due to sewage discharge from the touristic infrastructures. the ecological quality of lakes affected by eutrophication can be best described by their phytoplankton, macrophyte or phytobenthos communities (kelly et al. 2016). in the case of phytoplankton, different studies used sampling methods leading to results impossible to compare. the more recent studies, which used sampling by bottles and the utermohl (1952) method, showed that the most abundant phytoplankton species are smallsize cyanobacteria, such as woronichinia naegeliana, microcystis aeruginosa or aphanocapsa delicatissima. these species were lost from the samples in the older studies, which used nets to collect phytoplankton. furthermore, they were not considered by musacchio (1981-1982) and marano and d’aprile (1991), who neglected in the algal counts the so-called “nanoplankton”, i.e. the smaller cells. fig. 9. total phytoplankton abundance between 1981 and 2015 in lago grande from january to december. no nco mm er cia l u se on ly literature record of monticchio lakes 53 on the contrary, data from the last two studies (ceccanti et al., 2007; spicciarelli and mirauda, 2015) are fully comparable and show a relevant seasonal pattern in phytoplankton abundance, with large differences between monthly values (fig. 9). this pattern leads to the conclusion that, even using the same methods for sampling and analysis, comparing phytoplankton abundance in different years is only possible if the full seasonal pattern has been covered, possibly with monthly, or even more frequent, sampling. macrophytes have also been frequently used to assess lake ecological status. for example, pall and moser (2009) proposed a multi-metric index based on vegetation density, macrophyte maximum growing depth (mmgd), presence of characteristic zonation, abundance of more or less tolerant species and similarity of species composition to that found in “reference”, pristine lakes. in particular, mmgd has been shown to be a lake trophic indicator. for example, on the basis of a large data set including 757 european lakes, søndergaard et al. (2013) investigated the ability of mmgd to be used for evaluating lake trophic status. they found a good correlation between mmgd and secchi disc depth, but weaker correlation between mmgd and nutrient concentration. in spite of its considerable year-to-year variability, they suggest that mmgd can be used as a trophic indicator for clear water lakes deeper than 4-5 m, such as monticchio lakes. the difficulty in finding a good correlation between nutrients and mmgd may be due do the large heterogeneity of the used data set. in fact, evaluating a long time series of mmgd data from a single lake, may and carvalho (2010) found that changes in mmgd in loch leven (scotland) reflected changes in phosphorus load from its catchment. furthermore, relevant increases in mmgd during lake recovery from eutrophication has also been documented (hint et al., 2013). historical record of mmgd dating back to late 1800 or early 1900 are available for a number of lakes around the world (spence, 1982), including both lakes monticchio (forti and trotter 1908). they can then be used to infer reference conditions for these lakes, and to evaluate their trophic evolution, avoiding the difficulty in comparing secchi disk depth related to the strong seasonal variability of this metric, presenting high values in winter when algal population are less abundant, as shown by the high value in february 2005 reported in tab. 1. using this approach, may and carvalho (2010) related mmgd to the nutrient input in loch leven (scotland), inferring the history of its eutrophication and re-oligotrophication and showing that mmgd in 2006 was close to the value measured in 1905, indicating almost complete lake recovery. in other lakes, old studies on macrophyte did not report mmgd. for example, nimemeier and hubert (1986) reviewed changes in macrophyte community due to eutrophication in clear lake (iowa, usa), reporting a decrease in species number and a shift from submerged to emergent species, but they estimated a reduction of the photic zone from 4 m in 1951 to 0.8-1.5 m in 1981 on the basis of the secchi disk depth. in this international context, and as in loch leven, in the monticchio lakes a comparison between macrophyte data collected in the early 1900s and in recent times was possible. similar data are probably available in old records for other lakes that can allow to extend this exercise, giving information on their history. conclusions an accurate analysis of the studies carried out on the monticchio lakes allowed us to show the effect of the strong human impact affecting them in the second half of the 20th century. their biological response was very different, because of the difference in water chemistry driven by their different morphology and hydrology. in lg, nutrient input leaded to algal blooms, cyanobacterial dominance, low water transparency and a strong decrease in mmgd, while in lp the relevant input of fe2+ by underground springs (see table 2) caused p concentration to stay very low, leading to low algal biomass and in increase in water transparency and mmgd. this paper underlines the importance and the limits of evaluating old studies and grey literature to infer ecosystem history and reference condition. among the limnological parameters, mmgd resulted the more reliable indicator of trophic status, being relatively independent from sampling methods and seasonal pattern. phytoplankton abundance and composition would have been a valuable indicator, if a standard protocol have been followed, and frequent (e.g., monthly) samples collected. irregular studies carried out with different and noncomparable approaches, such those carried out in lake monticchio, may give important hints about lake history, but they cannot replace regular sampling with well-defined protocols, such those used in the long-term ecological research (lter) network. references alfonso g, 2008. [composizione e struttura dello zooplancton nei laghi dell’appennino meridionale].[phd thesis in italian]. università degli studi del salento, lecce. allen jrm, brandt u, brauer a, hubberten hw, huntley b, keller j, kraml m, mackensen a, mingram j, negendank jfw, nowaczyk nr, oberhänsli h, watts wa, wulf s, zolitschka b, 1999. rapid environmental changes in southern europe during the last glacial period. nature 400:740-743. azzella mm, iberite m, blasi c, 2010. [flora, vegetazione e inno nco mm er cia l u se on ly r. spicciarelli and a. marchetto54 dicatori macrofitici dei laghi vulcanici d’italia], p. 225-239. [article in italian]. proceedings 19th congr. italian society of ecology, bozen; 2009. baldi e, tonolli v, tonolli pirocchi l, 1953. [la differente evoluzione di due laghi già costituenti un unico bacino: il lago maggiore e il lago di mergozzo].[article in italian]. mem. ist. ital. idrobiol. 7:49-107. brauer a, allen jrm, mingram j, dulski p, wulf s, huntley b, 2007. evidence for last interglacial chronology and environmental change from southern europe. proc. natl. acad. sci. usa 104:450-455. cannicci g, 1952. [osservazioni sullo zooplancton di alcuni laghi appenninici naturali e artificiali].[article in italian]. boll. zool. 19:327-347. caracausi a, nuccio pm, favara r, nicolosi m, paternoster m, 2009. gas hazard assessment at the monticchio crater lakes of mt. vulture, a volcano in southern italy. terra nova 21:83-87. caracausi a, nicolosi m, nuccio pm, favara r, paternoster m, rosciglione a, 2013. geochemical insight into differences in the physical structures and dynamics of two adjacent maar lakes at mt. vulture volcano (southern italy). geochem. geophys. geosyst. 14:1411-1434. caracausi a, paternoster m, nuccio pm, 2015. mantle co2 degassing at mt. vulture volcano (italy): relationship between co2 outgassing volcanoes and the time of their last eruption. earth planet. sc. lett. 411:268-280. carvalho l, solimini a, phillips g, van den berg m, pietilainen op, lyche solheim a, poikane s, mischke u, 2008. chlorophyll reference conditions for european lake types used for intercalibration of ecological status. aquat. ecol. 42:203-211. casoria e, 1901. [le acque carboniche delle falde orientali del vulture].[book in italian]. premiato stab. tip. vesuviano, portici: 40 pp. cavanna g, 1882. [al vulture ed al pollino. narrazione della escursione fatta al vulture ed al pollino nel luglio del 1880 da a. biondi, c. caroti e g. cavanna].[article in italian]. boll. soc. entomol. it. 14:3-87. ceccanti b, de bernardi r, guilizzoni p, lami a, lepore l, marchetto a, masciandaro g, morabito g, mosello r, panzani p, poggio g, sili c, tornimbeni o, garibaldi l, leoni b, varallo a, 2007. [relazione sullo stato ambientale del lago grande di monticchio e sugli interventi di risanamento necessari].[report in italian]. report 04.07. cnr-ise, verbania: 90 pp. celico p, summa g, 2004. [idrogeologia dell’area del vulture (basilicata)].[article in italian]. boll. soc. geol. it. 123: 343-356. chiodini g, cioni r, guidi m, marini l, principe c, raco b, 1997. water and gas chemistry of the lake piccolo of monticchio (mt. vulture, italy). current res. volcanic lakes 10:3-8. chiodini g, frondini f, cardellini c, parello f, peruzzi l, 2000. rate of diffuse carbon dioxide earth degassing estimated from carbon balance of regional aquifers: the case of central apennine, italy. j. geophys. res. 105:8423-8434. cioni, r, marini l, raco b, 2006. [il lago piccolo di monticchio: geochimica dei fluidi e valutazione del rischio di eruzione limnica], p. 171-177. in: c. principe (ed.) [la geologia del monte vulture].[book in italian]. regione basilicata, potenza. d’ancona u, mozzi c, merlo s, 1961. [ricerche limnologiche sul lago di garda].[article in italian]. verh. int. ver. limnol. 14:838-845. de giorgi c, 1879. [note geologiche sulla basilicata].[book in italian]. tipografia salentina, lecce: 152 pp. de lorenzo g, 1900. [studio geologico del monte vulture].[book in italian]. atti r. accademia scienze fisiche e matematiche, naples: 207 pp. dumontet s, scopa a, pasquale v, franzese pp, melchiorre r, attolico a, allegretti g, 2003. [i laghi di monticchio e la fiumara di atella: indagine chimico-microbiologica e g.i.s].[report in italian]. ambiente e territorio, quaderno n. 4, supplemento. provincia di potenza. forti a, trotter a, 1908. [materiali per una monografia limnologica dei laghi craterici del m. vùlture. ann. bot., tipografia voghera, roma, vii, supplemento, 111 pp. gambardella b, 2006. flussi di co2 profonda disciolta nelle acque sotterranee del monte vulture], p. 155-169. in: c. principe (ed.) [la geologia del monte vulture].[book in italian]. regione basilicata, potenza. giannandrea p, principe c, la volpe l, schiattarella m, 2006. mappa geologica 1:25 000 del monte vulture].[map in italian]. in: c. principe (ed.) the geology of mount vulture, regione basilicata. consiglio nazionale delle ricerche. guilizzoni p, marchetto a, lami a, gerli s, musazzi s, 2011. use of sedimentary pigments to infer past phosphorus concentration in lakes. j. paleolimnol. 45:33-445. gussone g, tenore m, 1843. [ragguaglio delle peregrinazioni effettuate nella state del 1838 in alcuni luoghi delle provincie di principato citeriore e di basilicata (letto alla reale accademia di scienze (napoli), nelle tornate del 10 dicembre 1839 e 14 gennaio 1840)] atti reale accademia delle scienze, sezione della società reale borbonica. stamperia reale, naples 5 335-452. hartig f, 1963. [per la prima volta una bramaea [sic] in europa].[article in italian]. boll. assoc. romana entomol. 18:5-7. häse c, gaedke u, seifried a, beese b, tilzer mm, 1998. phytoplankton response to re-oligotrophication in large and deep lake constance: photosynthetic rates and chlorophyll concentrations. adv. limnol. 53:159-178. hilt s, kohler j, adrian r, monaghan mt, 2013. clear, crashing, turbid and back – long-term changes in macrophyte assemblages in a shallow lake. freshwater biol. 58:2027-2036. keating ej, dodd va, 1975. eutrophication of an inland lake in ireland in association with the intensification of pig farming in the catchment areas. agricult. environ. 2:55-64. kelly mg, birk s, willby nj, denys l, drakare s, kahlert m, karjalainen sm, marchetto a, pitt j-a, urbanič g, poikane s, 2016. redundancy in the ecological assessment of lakes: are phytoplankton, macrophytes and phytobenthos all necessary? sci. total environ. 568:594-602. la volpe l, patella d, rapisardi l, tramacere a, 1984. the evolution of the monte vulture volcano (southern italy): inferences from volcanological, geological and deep dipole electrical soundings data. j. volcanol. geothermal. res. 22 147-162. leoni b, garibaldi l, 2009. population dynamics of chaoborus flavicans and daphnia spp.: effects on a zooplankton community in a volcanic eutrophic lake with naturally high no nco mm er cia l u se on ly literature record of monticchio lakes 55 metal concentrations (l. monticchio grande, southern italy). j. limnol. 68:37-45. malpica c, 1847. [la basilicata: impressioni].[book in italian]. tipografia a. festa, naples: 255 pp. marano g, d’aprile s, 1991. [indagine sull’ecosistema dei laghi di monticchio].[report in italian]. consorzio turistico monticchio p.i.m. basilicata, sottoprogramma n. 2, bari: 94 pp. marchetto a, 1998. the study of high mountain lakes in the activity of the istituto italiano di idrobiologia. mem. ist. ital. idrobiol. 57:1-10. margaritora fg, 1992. limnology in latium: the volcan lakes. mem. ist. ital. idrobiol. 50:319-336. marinelli o, 1895. [area e profondità dei principali laghi italiani].[article in italian]. rivista geografica italiana, firenze. may l, carvalho l, 2010. maximum growing depth of macrophytes in loch leven, scotland, united kingdom, in relation to historical changes in estimated phosphorus loading. hydrobiologia 646:123-131. minder l, 1938. [der ziirichsee als eutrophierungs-phenomen. summerische ergebnisse aus fiinfzig jahren ziirichseeforschung].[article in german]. geol. meere binnengewasser. 2:284-299. mongelli f, 1964. [influenza delle acque sotterranee sul regime termico dell’apparato vulcanico del vulture].[article in italian]. boll. geofis. teor. appl. 6:24. mongelli f, panichi c, tongiorgi e, 1975. [studio termico ed isotopico dei crateri-laghi di monticchio (lucania)].[ article in italian]. arch. oceanogr. limnol. 18:167-188. morabito g, oggioni a, caravati e, 2005. decadal trends of pelagic algal biomass capacities in lago maggiore (n. italy). verh. internat. verein. limnol. 29:231-234. morabito g, mazzocchi mg, salmaso n, zingone a, bergami c, flaim g, accoroni s, basset a, bastianini m, belmonte g, bernardi aubry f, bertani i, bresciani m, buzzi f, cabrini m, camatti e, caroppo c, cataletto b, castellano m, del negro p, de olazabal a, di capua i, elia ac, fornasaro d, giallain m, grilli f, leoni b, lipizer m, longobardi l, ludovisi a, lugliè a, manca m, margiotta f, mariani ma, marini m, marzocchi m, obertegger u, oggioni a, padedda bm, pansera m, piscia r, povero p, pulina s, romagnoli t, rosati i, rossetti g, rubino f, sarno d, satta ct, sechi n, stanca e, tirelli v, totti c, pugnetti a, 2018. plankton dynamics across the freshwater, transitional and marine research sites of the lter-italy network. patterns, fluctuations, drivers. sci .total environ. 627:373-387. moretti g, 1958. [il lago trasimeno (tre anni di studi idrobiologici)].[article in italian]. quaderni sez perugina soc it biol sper 21:1-230. musacchio a, 1981-1982. [il fitoplancton invernale ed estivo dei principali bacini della basilicata].[article in italian]. delpinoa 23-24:99-113. nicolosi m, 2010. the monticchio crater lakes: fluid geochemistry and circulation dynamics. phd thesis, university of palermo. niemeier pe, hubert wa, 1986. the 85-year history of the aquatic macrophyte species composition in a eutrophic prairie lake (united-states). aquat. bot. 25:83-89. paci gm, 1853. [relazione dei tremuoti di basilicata del 1851 pel dottor giacomo m. paci].[report in italian]. regio ministero dell’interno, rome. pall k, moser v, 2009. austrian index macrophytes (aim-module 1) for lakes: a water framework directive compliant assessment system for lakes using aquatic macrophytes. hydrobiologia 633:83-104. palmieri l, scacchi a, 1852. [della regione vulcanica del monte vulture e del tremuoto ivi avvenuto nel di 14 agosto 1851: relazione fatta per incarico della reale accademia delle scienze].[book in italian]. stabilimento tipografico gaetano nobile, naples: 160 pp. poikane s, birk s, böhmer j, carvalho l, de hoyos c, gassner h, hellsten s, kelly m, lyche solheim a, olin m, pall k, phillips g, portielje r, ritterbusch d, sandin l, schartau a-k, solimini ag, van den berg m, van de bund w, 2015. a hitchhiker’s guide to european lake ecological assessment and intercalibration. ecol. indic. 52:533-544. poma g, salerno f, roscioli c, novati s, guzzella l, 2017. persistent organic pollutants in sediments of high-altitude alpine ponds within stelvio national park, italian alps. inland waters 7:34-44. principe c, 2006. [geologia del monte vulture. cnr, regione basilicata (italy)].[book in italian]. grafiche finiguerra, lavello: 217 pp. rogora m, massaferro j, marchetto a, tartari g, mosello r, 2008. the water chemistry of some shallow lakes in northern patagonian and their nitrogen status in comparison with remote lakes in different regions of the globe. j limnol 67:75-86. ruffo s, stoch f, 2005. [checklist e distribuzione della fauna italiana].[book in italian]. memorie museo civico storia naturale, verona: 307 pp. schettler g, albéric p, 2008. laghi di monticchio (southern italy, region basilicata): genesis of sediments a geochemical study. j paleolimnol 40:529-556. søndergaard m, phillips g, hellsten s, kolada a, ecke f, mäemets h, mjelde m, azzella mm, oggioni a, 2013. maximum growing depth of submerged macrophytes in european lakes. hydrobiologia 704:165-177. spence dhn, 1982. the zonation of plants in freshwater lakes. adv. ecol. res. 12:37-125. spicciarelli r, 2004. [la psiche del frassino].[book in italian]. consiglio regionale della basilicata, finiguerra arti grafiche, lavello: 176 pp. spicciarelli r, 2013. [la riserva naturale statale orientata “grotticelle”].[book in italian]. ministero delle politiche agricole alimentari e forestali, corpo forestale dello stato, stes, potenza: 157 pp. spicciarelli r, 2015. il vulture], p. 107-134. in: v. giacanelli, r. guarino, p. menegoni and s. pignatti (eds.), [sistemi ambientali e rete natura 2000 della regione basilicata, montagne e complessi vulcanici].[book in italian]. dipartimento ambiente e territorio, infrastrutture, opere pubbliche e trasporti, regione basilicata. grafiche zaccara, lagonegro. spicciarelli r, 2018a. bioethological observations on brahmaea (acanthobrahmaea) europaea (lepidoptera: brahmaeidae) in its dedicated nature reserve of “grotticelle” (basilicata, southern italy). in: a. zilli (ed.), lepidoptera research in areas with high biodiversity potential in italy. natura edizioni scientifiche, bologna. no nco mm er cia l u se on ly r. spicciarelli and a. marchetto56 spicciarelli r, 2018b. new bioethological observations on brahmaea (acanthobrahmaea) europaea and its host plants in special area of conservation “grotticelle di monticchio” (basilicata, southern italy) (lepidoptera: brahmaeidae). fragmenta entomol. 50:43-52. doi: https://doi.org/10.4081/fe.2018.283 spicciarelli r, mirauda d, 2015. [misure intese a preservare e sviluppare la fauna e la flora acquatiche presenti nel lago grande di monticchio (rete natura 2000, z.p.s. “monte vulture” it9210210)” ].[report in italian]. ue regione basilicata po-fep 2007-2013 misura 3.2., ees-reos, potenza. spicciarelli r, rosati l, mirauda d, azzella m, potenza g, fascetti s, 2016. [attuale distribuzione e potenziale invasività di taxodium distichum (l.) rich lungo le sponde del lago grande di monticchio (basilicata, italia meridionale)].[article in italian]. proceedings 11th nat. congr. on biodiversity, matera, italy. spicciarelli r, marchetto a, 2017. limnological studies of monticchio lakes (southern italy): a historical overview. proceedings 23th congr. italian association of oceanology and limnology, cagliari, italy. squicciarini m, 1974. [geomorfologia dei laghi di monticchio].[article in italian]. ist. geograf. univer. bari 1:1-33. stegagno g, 1908. [i crateri-laghi di monticchio (monte vulture)].[article in italian]. mondo sotterraneo, rivista di speleologia e idrologia. tip. d. del bianco, udine, vol. iv: 39 pp. stoddard jl, larsen dp, hawkins cp, johnson rk, norris rh, 2006. setting expectations for the ecological condition of streams: the concept of reference condition. ecol. applicat. 16:1267-1276. stoppa f, principe c, 1997. eruption style and petrology of a new carbonati tic sute from the mt. vulture southern italy: the monticchio lakes formation. j. volcanol. geothermal. res. 78:251-265. tadonléké rd, lazzarotto j, anneville o, druart jc, 2009. phytoplankton productivity increased in lake geneva despite phosphorus loading reduction. j. plankton res. 31:1179-1194. tata d, 1778. [lettera sul monte vulture a sua eccellenza il signor d. guglielmo hamilton dell’abate domenico tata].[book in italian]. stamperia simoniana, naples: 233 pp. tenore m, 1844.[ sul ciprino del vulture. atti della reale accademia delle scienze. memoria “letta nella tornata del 4 settembre 1838”].[article in italian]. vol. v, parte2a, stamperia reale, naples. teubner k, dokulil mt, 2002. ecological stoichiometry of tn:tp:srsi in freshwaters: nutrient ratios and seasonal shifts in phytoplankton assemblages. arch. hydrobiol. 154:625-646. tortorella f, 1840. [ragguaglio, osservazioni, e congetture fatte da paolino tortorella su di un fenomeno accaduto a’ 31 maggio e 31 luglio del 1810 ne’ laghi del monte vulture, volgarmente detti laghi di monticchio].[article in italian]. giornale economico letterario della basilicata, 1(ii):62-69. venanzoni r, apruzzese a, gigante d, suanno g, vale f, 2003. [contributo alla conoscenza della vegetazione acquatica e igrofitica dei laghi di monticchio].[article in italian]. inf. bot. it. firenze 35:69-80. vinciguerra d, 1895. [dell’opportunità di estendere gli studi limnologici a tutti i laghi italiani e dei metodi con cui condurli].[article in italian]. proceedings 2nd italian geography congr. g. civelli editore, rome. vollenweider ra, munawar m, stadelma p, 1974. comparative review of phytoplankton and primary production in laurentian great lakes. j fish. res. board can. 31:739-762 wallin m, wiederholm t, johnson rk, 2003. guidance on establishing reference conditions and ecological status class boundaries for inland surface waters. cis working group 2.3 – refcond. 2003-03-05: 93 pp. watts wa, 1985. a long pollen record from laghi di monticchio, southern italy: a preliminary account. j. geol. soc. london 142:491-499. wulf s, keller j, paterne m, mingram j, lauterbach s, opitz s, sottili g, giaccio b, albert pg, satow c, tomlinson el, viccaro m, brauer a, 2012. the 100-133 ka record of italian explosive volcanism and revised tephrochronology of lago grande di monticchio. quaternary sci. rev. 58:104-123. wulf s, kraml m, brauer a, keller j, negendank jfw, 2004. tephrochronology of the 100 ka lacustrine sediment record of lago grande di monticchio (southern italy). quatern. internat. 122:7-30. wulf s, kraml m, keller j, 2008. towards a detailed distal tephrostratigraphy in the central mediterranean: the last 20,000 yrs record of lago grande di monticchio. j. volcanol. geothermal. res. 177:118-132. corresponding author: renato.spicciarelli@unibas.it key words: historical data; water chemistry; phytoplankton; zooplankton; macrophytes; eutrophication. received: 24 november 2018. accepted: 18 june 2019. this work is licensed under a creative commons attribution noncommercial 4.0 license (cc by-nc 4.0). ©copyright: the author(s), 2019 licensee pagepress, italy advances in oceanography and limnology, 2019; 10:7949 doi: 10.4081/aiol.2019.7949 no nco mm er cia l u se on ly layout 1 introduction the continental shelves are characterised by large amounts of suspended particulate matter, derived from various sources as the autochthonous production, rivers and terrestrial runoff. the northern adriatic sea (nad) shelf receives large riverine inputs of dissolved organic carbon, nutrients, suspended solids and associated carbonates, organic carbon, nitrogen and phosphorus (pettine et al., 1998; ludwig et al., 2009; cozzi and giani, 2011; cozzi et al., 2012). as a result, the nad is one of the most productive sectors of the entire mediterranean sea (d’ortenzio and ribera, 2009; lazzari et al., 2012). in the last decade, a shift toward a general oligotrophication of the nad has been reported (mozetič et al., 2010; giani et al., 2012; colella et al., 2016) along with an increasing trend of water acidification (luchetta et al., 2010). since these phenomena are expected to exacerbate in the near future and expand down to the sea bottom, a better comprehension of the benthic-pelagic coupling processes in the nad and, in particular, the quantification of the downward particle fluxes is crucial for understanding ecosystem changes in this shallow continental shelf. the concentration and composition of suspended particulate matter and the related downward fluxes in the nad generally show a large spatial and temporal variability (giani et al., 2001). the typical environmental conditions (e.g., shallow depth not exceeding 50 m, large riverine inputs, cyclonic circulation) and the seasonal cycle determine the persistence of rather different biogeochemical processes in the water column and at the water-sediment interface (ogrinc et al., 2003; solidoro et al., 2009). the most common methodology for estimating the downward flux of suspended particulate matter uses sediment traps. even if the sediment traps deployed in shallow waters are subject to different bias due to hydrodynamic effects of the flow over the trap mouth and to the presence of fouling and swimmers (buessler et al., 2007), these instruments remain a solid tool to date for advances in oceanography and limnology, 2018; 9(2): 45-58 article doi: 10.4081/aiol.2018.7601 this work is licensed under a creative commons attribution-noncommercial 4.0 international license (cc by-nc 4.0). fluxes of particulate matter, carbonates, organic carbon and nitrogen in the northern adriatic continental shelf: a synthesis overview michele giani,1* juan carlos miquel,2# amelia de lazzari,3 alfredo boldrin3 1istituto nazionale di oceanografia e di geofisica sperimentale (ogs), via a. piccard 54, 34151 trieste, italy; 2environment laboratories, international atomic energy agency, monaco; 3istituto di scienze marine ismar-cnr castello 2737/f, 30122 venezia, italy #present address: institute bobby, cap d’ail, france *corresponding author: mgiani@inogs.it abstract time series of composition and fluxes of settling particles in the marine environment, obtained by sediment traps, contribute to define the main processes driving the dynamics of particulate matter and of the time/space variability of benthic-pelagic exchanges. with this aim, the composition and seasonal and annual fluxes of settling matter, obtained from different projects and from published papers, at 13 sites of the northern adriatic shelf were estimated. the mean yearly particulate fluxes varied from 2763 to 14447 g m–2 y–1, from 66 to 236 gc m–2y–1 for organic carbon (oc) flux, from 861 to 7525 g m–2y–1 for carbonates and from 12 to 42 gn m–2y–1 for nitrogen (n). the fluxes were characterized by high seasonal variations with marked increase in autumn or in winter with respect to spring or summer. the sink of particles occurs in relatively short episodes as about 50% of annual particle flux settles in less than 1-2 months in the western coastal area. this seasonality can be related to the riverine discharges, primary production and wind regimes of the basin. utilizing the n/oc ratio as an index for discriminating the different origin of organic matter (i.e., resuspended/riverine and autochthonous), the primary marine carbon flux was estimated to range from 10 to 28% of the oc fluxes and accounted for 8-40% of the primary production, depending on the site. then, due to the shallow waters of the basin and to the relevant riverine inputs, the total fluxes near the sea bottom were highly dependent on resuspension and advective transport processes. the important contribution of these last processes as source of organic matter is suggested also by the comparison between fluxes determined by sediment traps with mass accumulation rates (mar) in sediments, derived from radionuclide measurements. indeed, the fraction of oc fluxes which is not buried in the sediment is sufficient to support the benthic respiration processes. key words: organic carbon; carbonates; nitrogen; settling particulate matter; adriatic sea; continental shelf. received: june 2018. accepted: november 2018. no nco mm er cia l u se on ly m. giani et al.46 estimating the particle flux from the water column to the sea floor and its temporal variability. due to the cyclonic circulation and to the geographic setting of river mouths, decreasing west-east and coastoffshore gradients in suspended matter concentrations, particulate organic carbon concentrations and plankton primary production are generally observed in the nad (gilmartin and revelante, 1983; smodlaka, 1986, giani et al., 2005). thus, our analyses considered the data obtained from a selection of sites located in three different sectors of the northern adriatic sea: western coastal, eastern coastal, and offshore sectors (fig. 1). the variables used for the analysis included the total mass flux as well as carbon, nitrogen and carbonates fluxes. we searched for and collated a large data set of downward particle fluxes measured using sediment traps in the nad. these data, extracted from published and grey literature and partly still unpublished, were analysed to ascertain spatio-temporal scales of variability in the benthic-pelagic exchanges of this shallow area of the adriatic sea. in shallow basins such as the nad the suspended particle dynamics typically depends not only upon the allochthonous inputs, local primary production and degradation/remineralization processes, but are also largely influenced by sediment resuspension and near-bottom advective transport induced by winds and waves (faganeli, 1989; puškaric et al., 1992; giani et al., 2001; boldrin et al., 2009). therefore, in this study the downward fluxes of organic carbon were compared with primary production in the different areas, with an estimation of primary marine fluxes of organic carbon in this basin. to evaluate the fraction of the settling particulate organic matter, which accumulates on the seabed, the particle flux was compared with the mass accumulation rates (mar), allowing to estimate the organic carbon potentially available for consumption by benthic organisms or chemical oxidation. methods study area the nad is subject to strong seasonal variations: in winter, because of pronounced vertical instability of the water column associated with seasonal wind mixing, po river freshwaters are mostly confined along the nad western coast, whereas during the period of stratification the po river plume can extend eastwards toward the istria peninsula (franco et al., 1992) as response to the intensity of the river discharge and to the bora wind (mauri et al., 2007). in the northern adriatic sea, major primary production variations occur along the trophic gradient due to nutrient rich freshwater discharges. the range of phytoplankton production is considerable: annual values range between 60 and 90 g c m−2 yr−1 and between 130 and 210 g c m−2 yr−1 for offshore and coastal waters, respectively (pugnetti et al., 2006 and references therein). fluvial sediments are initially deposited on the prodelta areas with an efficient transport along the shelf (coreggiari et al., 2001), where fine sediments are deposited mainly in a belt parallel to the western coast which get wider south of the po prodelta (brambati et al., fig. 1. location of sediment traps deployment. triangles indicate moored trap positions, whereas the circles represent the central position of drifting trap deployments. no nco mm er cia l u se on ly fluxes of particulate matter, carbonates, organic carbon and nitrogen in the northern adriatic continental shelf 47 1983). in nad shallow waters the strong hydrodynamic forcing, mainly determined by ne and se winds, can promote sediment resuspension (wang and pinardi, 2002; fox et al., 2004; fain et al., 2007; boldrin et al., 2009). mass accumulation rates in the nad range from 0.06 to 6.6 g cm–2 y–1 (corresponding to 0.6 to 66 kg m–2 y–1) with the highest values observed in the po and isonzo river prodeltas (frignani et al., 2005). site location and sediment traps the sediment trap data used in this work was collected in the framework of different research projects spanning from 1982 to 2010. we considered 10 sites with moored trap and 3 with drifting traps located in the nad sites shown in fig. 1, the metadata describing the trap deployments at the investigated sites are reported in tab. 1. in all the sites, the moored and the drifting traps were sited few meters above the bottom (bottom-traps). in 5 sites for moored and 2 for drifting traps an additional upper trap was positioned in the middle of the water column. part of sediment-trap data have been previously published (faganeli, 1989; puškaric et al., 1992; miquel et al., 1999; giani et al., 2001; turchetto et al., 2002), but here they have been reconsidered, re-elaborated and further discussed. other data, collected by the authors but only partially published (boldrin et al., 2006; giani et al., 2003b) or unpublished (boldrin and de lazzari, personal commumincation), have been utilised in this work. the sites with moored traps can be geographically divided in western (sites pa, s1, s2, po, cs), eastern (sites mi, pi, pu) and offshore (sites ro, s3) areas. whereas, the drifting traps were located in western (site a) and offshore (sites b, c) areas. a total of 374 moored trap samples were considered (327 for bottom-traps and 47 for upper-traps), 219 in western, 73 in the eastern and 82 in offshore areas. overall, the sampled periods amounted to 3710 days for bottom-traps (western: 1598; eastern: 1588; offshore: 524 tab. 1. site geographic settings and sediment trap general information. no nco mm er cia l u se on ly m. giani et al.48 days) and to 1104 days for the upper-traps (split up in: western 231, eastern 175 and offshore 698 days). the western sites s1 and po are located in front of the po river delta, whereas s2 is influenced by adige river. these rivers represent the main freshwater loads in the nad (average discharges 1569 m3 s–1 and 200 m3 s–1, respectively; cozzi and giani, 2011). minor rivers (sile and piave: 41 m3 s–1 and 54 m3 s–1 average discharges, respectively; autorità di bacino dell’adige, 2009a and 2009b, arpav, 2018, verri et al., 2018) influence the cs site. the site pa, located at about 15 km from venetian coast, is considered in the western area. the easternmost site (mi), in the gulf of trieste, is affected by the isonzo river discharge (average 82 m3 s–1; cozzi et al., 2012), only during high floods. moreover, in the eastern area we considered also other two sites (fig. 1), one located in front of pula (site pu; puškaric et al., 1992), and the second in the bay of piran, in the south-eastern part of trieste gulf (site pi), and reported earlier by faganeli (1989). the site s3, located at about 40 km east of venetian coast (giani et al., 2001), and the site ro, 25 km off rovinj (puškaric et al., 1992), are considered as representative of the offshore environment. furthermore, drifting sediment trap located inside the po river plume (trap a) and outside of it (traps b and c) were considered. these data were obtained from the studies carried out by turchetto and co-workers (2002) in june 1996, june 1997, february 1997, february 1998 and by miquel and coworkers (1999) in july 1993. more information on sites and details on deployments are reported in tab. 1 and the localisation is shown in the fig. 1. different designs of the traps were used in the deployments and could affect the efficiency of the sampling device, not easy to quantify (for a complete analysis of the biases in the sediment traps use see usgofs, 1989). however, all the traps used were cylindrical-conical type with height to width ratios (h/w) generally ≥4, required for good collection efficiency under most environmental conditions (buesseler et al., 2007). the sampling frequency varied from 1 to 91 days. the difference in sampling timing introduces potential biases due to the fouling of the traps during prolonged absence of maintenance, though the longer sampling concerned the traps located in the offshore waters. treatment of sediment trap samples the sample treatment and the analytical methods were similar in all the experiments. collection bottles were filled with 4% formaldehyde solution of seawater collected at the same depth of the traps, buffered at ph 8.2 and filtered on a 0.4 µm nuclepore filter (miquel et al., 1994). in the open cylindrical traps used in s3 and partially in pa, to minimize diffusive loss of poison during deployment and better retain the collected samples, the sampling bottles, screwed at the base of the tube, were filled with 4% formaldehyde solution of a 50 g l–1 nacl solution (usgofs, 1989). samples treatment followed the methodology described in heussner et al., 1990. the “swimmers”, metazoan zooplankton that actively enter sediment traps and that do not contribute to the passive flux (buesseler, 2007; rizzo et al., 2009; miquel et al., 2011), were removed manually under a dissecting microscope. to determine total mass flux (tmf) the gravimetric method was used, whereas organic carbon (oc), total carbon (tc) and total nitrogen (n) contents were determined using high temperature gas-chromatographic chn elemental analysers. organic carbon was measured after elimination of carbonates by acidification (hedges, 1984). inorganic carbon content was calculated as the difference between tc and oc; the carbonate content was calculated by assuming all inorganic carbon was caco3, using the ratio between molecular weights of caco3 and c of 8.33. sediment-trap data analysis the seasonal fluxes were calculated as time-weighted means. to obtain the time-weighted flux (fitw), each of individual flux measurements (fi) was weighted by the duration of collection (number of days di) as: fitw=σ(fi di)/σ(di). this time-weighted mean gives the best (i.e., the least biased) estimate for discontinuous measurements and is used for the determination of seasonal values. then the annual mean was calculated from the seasonal values multiplied for the season duration. the seasonal periods were defined as: winter=january-february-march; spring=april-may-june; summer=july-august-september; autumn=october-november-december. the comparison of differences between sites and seasons was performed using the kruskal wallis nonparametric test. the site s1 was excluded as the data for spring were lacking. the software statistical package statistica™, ver. 6 (statsoft®, usa) was used. additional data the flux data were integrated with available information on the composition of particulate matter and primary production of the water column and sediment properties of the sites. the carbon export was estimated as the ratio between the oc primary downward flux and the carbon produced by phytoplankton. annual primary production data were obtained by previous studies carried out at s1, s2 and s3 sites (giani et al., 2001), at cs site (boldrin et al., 2006) and in the gulf of trieste (fonda umani et al., 2007). the elemental composition of the top first cm of sediments (0-1 cm layer) at the several stations considered was obtained from the literature and final project reports no nco mm er cia l u se on ly fluxes of particulate matter, carbonates, organic carbon and nitrogen in the northern adriatic continental shelf 49 (faganeli et al., 1991; ogorelec et al., 1991; giani et al., 1997; de lazzari et al., 2006), whereas the data for site pa are from boldrin (unpublished data). the fraction of sinking carbon that definitely settles on the seabed was estimated on the basis of the carbon sedimentation rates (ocsed). in the absence of direct measurements in our experimental sites, we used the nearest mass accumulation rate (mar) value measured on the basis of activity–depth profiles of 210pb and reported in the literature by frignani et al. (2005) and the oc content determined in the sediments at the study sites. for the site ro the mass accumulation (mar) rates were estimated from sediment accumulation rates (sar) on the basis of a water content of 29.9% and a density of 2.69 g cm–3. results and discussion downward fluxes and composition of settling matter the averages of tmf and elemental composition of settling particulate matter at all the investigated sites are reported in tab. 2. overall, tmf data for bottom-traps ranged between 0.05 and 293.50 g m–2 d–1, with an average of 22.35±40.87 g m–2 d–1. the downward fluxes follow the spatial trend of suspended particulate matter concentrations typically observed in the nad (giani et al., 2003a; boldrin et al., 2005), with a clear (p<0.01, mann-whitney u test) increase of tmf in the coastal-western area with respect to the eastern coast and the offshore areas. the western coastal sites showed higher fluxes (mean value: 30.96±49.60 g m–2 d–1), whereas the eastern sites flux was on average reduced to about 1/3 (11.00±15.88 g m–2 d–1) and to 1/4 in the offshore sites (7.19±9.16 g m–2 d–1), but the differences between these last two areas were not significant. tmf in the upper-traps showed the same spatial trend of the bottom-traps, but the fluxes were about 5 time lower (average 4.63±5.79 g m–2 d–1; in the western area 7.25±7.24 g m–2 d–1; in the eastern sites 2.18±0.58 g m–2 d–1; offshore 2.09±2.13 g m–2 d–1). the drifting-trap mass fluxes were around 0.51±0.48 g m–2 d–1, about 3% of the moored bottom-ones in the western coastal area and about 6% in the offshore. this discrepancy tab. 2. daily total mass flux (tmf) and elemental composition of settling particles. no nco mm er cia l u se on ly m. giani et al.50 could be considered very great but we must consider the difference in the methodology, that could reduce the lateral supply and the difference in sampling period, as more widely discussed in buesseler et al. (2007). the poc fluxes in the drifting traps ranged from 1 to 82 mg m–2 d–1. the highest values are comparable to the highest ones reported for the mediterranean sea in the alboran sea (ramondnec et al., 2016). the mean particulate nitrogen fluxes were 3fold higher than those measured in the nw mediterranean sea (marty et al., 2009). the offshore fluxes of oc were 13 mg m–2 d–1, less than half of those measured by the more coastal traps, and more than 10-fold lower by the fluxes measured with the same approach in the gulf of trieste (wassmann et al., 1999). the oc and n contents in the settling matter of bottom traps ranged from 0.71 to 11.30% (average 3.01±1.57%) and from 0.06 to 2.21% (average 0.42±0.29%), respectively. in the upper-traps the percentages of oc and n on tmf were about double than in the bottom ones, being on average 6.14±3.63% for oc and 0.92±0.57% for n. at all sites, the oc content of settling particles was linearly correlated with that of n (p≤0.0001). in the uppertraps the oc/n molar ratio derived from the linear regression slope was 7.12, whilst in bottom-traps oc/n was 9.86. this difference could be attributable to a higher contribution of plankton in the upper trap than in the bottom one, where also resuspended matter can easily settle. carbonates, which constituted from 2 to 59% of the total mass flux, are provided by the minerals carried by rivers (calcite and dolomite), by calcareous planktonic organisms as coccolithophorids (turchetto et al., 2002; bernardi aubry et al., 2004) or foraminifera, and by resuspension of bottom sediments containing biogenic carbonates (calcite and aragonite) from shell and skeleton debris, coralline algae (ogorelec et al., 1991) and benthic foraminifera (puskaric et al., 1992). calcite and dolomite minerals are mostly carried in the nad by isonzo, piave, tagliamento, brenta and adige rivers (brondi et al., 1979), whereas the po river carries no dolomite, less calcite and higher amounts of silica particles. the higher content of carbonates (mann-whitney u test, p<0.05), found in areas influenced by adige and brenta rivers (site s2), by piave (site cs) and by the outlets of the venice lagoon (site pa), reflect the contribution of calcite and dolomite carried by the rivers with drainage basin including the dolomitic alps, as also shown by the distribution of carbonates in the surface sediments of the nad (ravaioli et al., 2003). high carbonate concentrations can be found also off the istria peninsula (ne adriatic sea) where the sediments contain more than 50% of carbonates originating from mesozoic sediments of dinaric alps (brambati et al., 1973). however, the elevated carbonate content in the trap material at pu site was likely due to a high recent biogenic contribution to the deposition of carbonates (puskaric et al., 1992). the data obtained from sediment trap located at the site mi apparently do not support an important role of the isonzo river as a carrier of carbonates. this incongruence is likely to be attributable to the fact that the site mi is located far from the prodelta, in a coastal zone where the carbonates in the sediments are less than 30% (ogorelec et al., 1991). furthermore, the hydrological circulation in the gulf of trieste is characterized, generally, by an inflow sw-ne current determining a cyclonic circulation at the sea bottom in all seasons (bogunović and malačič, 2009; malačič and petelin, 2009), preventing the direct supply of isonzo sediments to the trap-site. oc and n content (% of dry weight) were inversely and exponentially correlated with tmf (fig. 2 a,b; fig. s1). such a “dilution effect” of oc and n within the tmf, early reported for the site s1 (matteucci et al.,1997) and observed also in other areas (e.g., in ne-mediterranean by stavrakakis et al., 2000; in southern adriatic by langone et al., 2016; in antarctica by tesi et al., 2012), could be related to the increase of the mineral fraction. this could be the result of ocand n-depleted resuspended sediments captured by the traps, as well as/or of an advective transport of riverine particulate matter (pettine et al., 1998). on the other hand, carbonates increase with tmf at sites s1, s2 and pa, suggesting that the above-mentioned dilution effect is most likely due to carbonate rich particles collected by the traps. the oc/n ratio increases with tmf, according to a power fit of the type oc/n=a(tmf)b at sites s1, s2 and cs (fig. 2), possibly due to the resuspension of bottom sediments which have higher oc/n molar ratios (tab. 3). the fittings reported in fig. 2 are in some cases asymptotic, then y values tend to a given limit and this asymptotic limit was compared with the corresponding content of surficial sediments (tab. 3). the asymptotic values differed for less than 10% from the contents of carbonates whereas for oc and n (only exception site s1) the estimated contents in settling matter were generally from 1.5to 7-fold higher than those in the sediment (tab. 3). these differences could be due to the degradation of labile organic matter, taking place during sediment accumulation, with respect to a more conservative component as the carbonates. in the settling matter collected by the drifting traps the chemical composition showed a higher content of organic carbon (14.78±8.31% on average) and nitrogen (2.28±1.43% on average) with respect to the moored traps which had an oc content of 4.19±2.33% and n content of 0.63±0.41% (tab. 2). the drifting traps in offshore waters were less influenced by resuspended sediments and riverine particulate matter as shown by the lower oc/n ratio (7.6±0.9 on average). this can be related to the short deployment period and the calm weather required during the experiments, conditions which limit particle resuspension. these ratios fall in the no nco mm er cia l u se on ly fluxes of particulate matter, carbonates, organic carbon and nitrogen in the northern adriatic continental shelf 51 range of the data reported for the eastern mediterranean sea by ramondenc et al. (2016). the high variability of the carbonate content could be due to two opposing processes: a lower contribution of resuspended sediments and a higher contribution of the settling of coccolithoforids in offshore waters which can be important especially during winter (godrijan, 2018; viličić et al, 2009). seasonal and annual fluxes the seasonal settling fluxes for total particles, organic carbon, nitrogen and carbonates, calculated as timeweighted averages, are showed in fig. 3. we tested spatial and temporal differences in the fluxes, using the non-parametric anova-kruskal-wallis test, considering, separately, 2 factors: sites and seasons (tab. 4). results of these tests showed that the effect of the site is significant for all parameters except for oc flux, whereas the seasons affect all the dependent variables except the caco3 content. the highest fluxes occur, with a few exceptions for oc and n fluxes, in winter or autumn (fig. 3). in those sites, as s1 and s2, directly influenced by the greater rivers tmf in winter were up to 9.6-fold higher than in summer. a similar difference is observed also in the tab. 3. mean composition and oc/n molar ratio of surface sediments, sediment (sar) and mass accumulation (mar) rates. fig. 2. a) organic carbon and calcium carbonate content, and oc/n molar ratio relationships with total mass fluxes for the sites s1, s2, pa, cs, s3 and mi; the best fits are represented by red lines. b) organic carbon content (%), calcium carbonate content (%), and oc/n molar ratio relationships with total mass fluxes for the drifting traps at a, b and c sites; the best fits are represented by red lines. no nco mm er cia l u se on ly m. giani et al.52 organic carbon and nitrogen fluxes which were up to about 5.5 times higher in winter than in summer. to further assess the extent of seasonal variability of the investigated fluxes, the seasonality index (si), proposed by berger and wefer (1990) and lampitt and antia (1997), was calculated for sites s2, pa and ro, for all of which almost complete annual time series were available. si is defined as the time (in days) needed for the deposition of 50% of the total matter sinking in one year after the ranking of the data (fig. 4). si for tmf varied from 26 days in site s2, to 66 in pa and up to 111 days in ro, and from 27 to 137 days for the oc flux in s2 and ro sites (the data from pa are not sufficient to complete this calculation). according to the classification proposed by berger and wefer (1990), the si values allow us to define the export “strongly seasonal to pulsed” in western coastal area, since about 50% of the annual deposition settles within 1-2 months. this value is considered characteristic of areas were the sedimentation is related to events occurring at seasonal scale (riverine discharges, phytoplankton blooms) and occasional events (such as storms) strongly influence the fluxes of settling sediments in nad. in the more oligotrophic offshore area (site ro) the 50% of sedimentation occurred in about 4-5 months and the flux appears more constant during the year. the annual fluxes obtained from bottom traps and calculated from the seasonal weighted mean (tab. 5), ranged from 2763 to 14447 g m–2y–1 for total mass flux, from 66 to 236 g oc m–2y–1 for organic carbon, from 12 to 42 g n m–2 y–1 for nitrogen and from 861 to 7525 g m–2y–1 for calcium carbonates. the lowest fluxes were observed in the offshore area, whereas the highest occurred in the western sites located in front of the river mouths. the high fluxes measured in the western site cs could be due not only to the riverine particulate deposition but also to the presence of the artificial barriers close to the sampling site (zennaro et al., 2006). it is known that the artificial reef, fig. 3. time-weighted seasonal averages of settling fluxes for total particles (tmf), organic carbon (oc) and calcium carbonates (caco3) in bottom-traps. tab. 4. non-parametric anova-kruskal-wallis tests on bottom-trap fluxes and composition (%), considering 2 independent factors: sites and seasons. no nco mm er cia l u se on ly fluxes of particulate matter, carbonates, organic carbon and nitrogen in the northern adriatic continental shelf 53 modifying the hydrodynamic of the area, could increase the turbulence caused by storm events and enhance the resuspension processes (ambrose and anderson, 1990; falcão et al., 2009). as the cs station is shallower than the other western stations, the energy due to wind stress reaching the bottom is higher and, therefore, it generates elevated resuspended fluxes, as also evidenced by the higher values of the ocr/oc ratio (tab. 6). the higher annual carbonate fluxes in cs and s2 sites indicate a contribution of the riverine suspended matter discharged from the piave and from adige rivers, respectively, whose drainage basins include carbonate areas (ravaioli et al., 2003). fig. 4. seasonal index (si) for accumulated total mass flux (a) and oc flux (b) for sites s2 (period june 1995-march 996), pa (period june 1993-june 1994) and ro (period januarydecember 1989). the relative time values (in days) for 50% of accumulated fluxes for each parameter are reported. tab. 5. annual fluxes of particulate matter, organic carbon, carbonates and nitrogen in the bottom traps. where the seasonal data was lacking in the computation of the annual fluxes, the value was estimated as the average of the other available values. tab. 6. gross (oc), resupended+riverine (ocr) and primary (ocp) annual fluxes of organic carbon compared to primary production (pp) and organic carbon sedimentation (ocsed), calculated from mass accumulation rates. no nco mm er cia l u se on ly m. giani et al.54 the annual tmf estimated in the northern part of the trieste gulf (4693 g m–2 y–1, site mi) is close to the flux reported in the south-eastern part of the gulf (site pi) by faganeli (1989). more generally, the annual oc fluxes in the whole nad are up to two-three order of magnitude higher than those of other mediterranean areas. the oc fluxes at southern adriatic pit were estimated in 3.3 g c m–2 year–1 and at northern ionian in 2.4 g c m–2 year–1, both at 150 m depth (boldrin et al., 2002). low value (4.8 g c m–2 year–1) was observed also in ligurian sea at 80 m depth (miquel et al., 1994). the higher fluxes observed in nad are reasonably correlated with the high productivity of the basin (pugnetti et al., 2006) and with the presence of elevated riverine inputs (cozzi and giani, 2011). in addition, the shallow water of the basin (on average 35 m; lipizer et al., 2014) increases the potential effects of resuspension and sediment lateral transport. resuspension and riverine contributions as already pointed out, the resuspension and sediment transport processes assume high relevance on the particle vertical fluxes in the nad. these processes are related to the interaction between the general circulation of the water masses and the shallow bathymetry of the basin (fain et al., 2007). moreover, the wind stress can cause resuspension and southward transport of bottom sediments, as observed during the ne-bora wind events (wang et al., 2007; boldrin et al., 2009), or when elevated wave heights resuspend sediments at shallow sites (giani et al., 2001). then, the particles collected by the sediment-traps can derive from different processes and have different origin. end members mixing models based on oc/n ratio and the stable carbon isotopic composition (δ13coc) have been frequently used in the nad to identify the sources of organic matter in the suspended and sedimentary matter (faganeli et al., 1988; boldrin et al., 2005; tesi et al., 2007; giani et al., 2009). in this study, to estimate the fraction that could attributed to biological production (primary flux) and those due to resuspension and transport of bottom sediment (secondary sedimentation), a label approach based on two-end members mixing was applied, according to the methodology proposed by gasith (1975) and previously applied in the adriatic sea (faganeli et al., 1989; matteucci et al., 1997; giani et al., 2001). for this purpose, n/oc was selected to distinguish resuspended sediments and/or riverine settling matter from autochthonous settling organic matter of planktonic origin (perdue et al., 2007). organic matter derived from terrestrial and marsh vascular plants is typically depleted in nitrogen (n/oc<0.07), whereas marine phytoplankton, zooplankton and bacterioplankton are characterised by higher nitrogen content (n/oc >0.13) (goñi and hedges,1995; goñi et al., 2000). to estimate the resuspended plus the allochthonous organic matter, the n/oc end members were selected for each site as they lie in different sedimentary environments. as end member representative of resuspension we considered the n/oc lower ratio (0.02 percentile of available data for each sediment trap) measured in settling matter (n/oc end members ranged from 0.07 to 0.09). as end member representative of marine particulate matter we used the 98th percentile of n/oc ratios in particulate suspended matter (n/oc end members ranged from 0.17 to 0.24). then, the estimated resuspended or allochthonous matter transported by rives (ocr) on an annual basis (tab. 6) ranged from 72 to 90% of the total particulate organic carbon fluxes. therefore, as observed in previous studies (matteucci et al., 1997; giani et al., 2001), the dominant part of the fluxes in the nad can be attributed to resuspension processes or riverine inputs of allochthonus particulate organic matter. organic carbon export and carbon accumulation in the sediments the export of autochthonous-primary organic carbon (ocp), calculated as the difference between oc and ocr, was compared with the available data of primary production (tab. 6), and the exported fraction ranged from 8 to 40%. the higher export corresponded to the site with lowest primary production (sites cs, pi and s3), whereas the lowest occurred in the most productive areas near the po and adige rivers (i.e. sites s1 and s2, respectively; fig. 5). in these sites, even if the production is higher due to the removal of nutrient limitation, the export appears lower for the presence of elevated lateral advection and, probably, because of an intense degradation of the organic matter fig. 5. primary organic carbon exported toward the sediment with respect to primary production (ocp export, expressed as percentage) and fraction of settled organic carbon (ocsed, expressed as percentage) with respect to the gross organic carbon flux measured in the bottom traps. no nco mm er cia l u se on ly fluxes of particulate matter, carbonates, organic carbon and nitrogen in the northern adriatic continental shelf 55 due to high bacterial activities (puddu et al., 1998; giani et al., 1999). moreover, the data and satellite images (brando et al., 2015) indicate that the influence of the riverine input on site cs is lower than in s1 and s2, and this could be likely due to the different discharges of rivers: po>> adige> piave > sile. the organic carbon settling fluxes, as determined with the bottom sediment traps, were from 2 to 39-fold higher than the organic carbon ones, calculated from mass accumulation rates at all the sites (tab. 6). this implies that, at all sites, a relevant fraction of the settling organic matter is not buried, rather is consumed by benthic organisms, decomposed or advected to other areas. as shown in fig. 5, the highest % of the oc flux settles in the sediments near the po river (site s1) and adige river (site s2). the mass accumulation rate (mar) integrate the sedimentation over years and, therefore, when it is converted in flux of organic carbon, it does not reflect the deposition of the labile fraction of organic matter, which could be respired or consumed by benthic organisms on a time scale of weeks or days (giordani et al., 2002). comparing the oc flux excess (oc-ocsed) with the annual carbon mineralized at the sediment water – interface in the western coastal nad, which falls in the range 54-89 gc m–2y–1 (moodley et al., 1998), we can estimate that from 25 to 78% of the organic carbon input to the sea bottom could be lost through benthic respiration at western coastal site under riverine influence (s1, s2, cs). if we take the respired carbon benthic fluxes estimated by giordani et al. (2002) for the western nad (100-130 gc m–2y–1) the incidence of the utilization of the total organic carbon particle flux rises up to 47-114%. notwithstanding the variation of these estimates, these results suggests that a quite relevant fraction of the settling oc which is not buried in the sediments is likely consumed or remineralized at the sediment-water interface. at the sites s1, s2 and cs, highly influenced by rivers, a significant fraction of oc accumulated in the sediments results from the riverine discharge or from lateral advection. therefore, we confirm that riverine or laterally advected labile organic matter must play a relevant role in sustaining the benthic trophic chain (danovaro et al., 2000). the particulate organic matter carried by the po river, especially during low discharge periods, is rich in phytoplankton (pettine et al., 1998) which presumably is a source of allochthonous labile organic matter. as a matter of fact, both bacteria and meiofauna were shown to respond to frontal areas created by plumes and their associated gradients of settled organic matter particles south of the po prodelta (danovaro et al., 2000). conclusions the downward fluxes of particles in the nad present strong gradients decreasing from west to east. the coastal areas influenced by higher riverine inputs (po and adige rivers) are characterised by higher productivity but lower carbon export when expressed as percentage of primary production. in offshore waters, primary production is lower but the organic carbon transfer to the bottom in percentage is about double with respect to the coastal areas. downward fluxes and the elemental composition of settling matter are characterised by high seasonal variations. the sink of particles occurs in relatively short episodes as about 50% of annual flux occurs in less than 2-3 months. in the shallow continental shelf of the northern adriatic, resuspension and advective transport, accounting for up to 70-90% of the total organic carbon flux at annual scale, are the main processes regulating the dynamic of particles, their concentration and sedimentation. the importance of these processes is highlighted by the relevance of carbonates, which represent a substantial component of the fluxes in the north-western area of the nad and are associated with the river inputs and with the dispersion of sediments through resuspension. at most sites, the autochthonous organic carbon fluxes were higher than the organic carbon accumulation at the surficial sediment as estimated by radionuclide data, and this excess likely contributes to sustain the biological and chemical processes in the sediment. in the areas influenced by higher riverine discharges, the total sedimentation of oc, estimated at annual scale, is much higher than the flux due to the primary production, then a relevant allochthonous source of organic matter is required to sustain the benthic community. future studies with sediment traps in the area should include a more detailed characterization of the lithogenic fraction and should also trace the organic matter sources with stable isotopes in order to better estimate the contributions of resuspension processes, and of riverine versus marine matter. acknowledgments part of this work was financially supported by vector-fisr project of the italian ministry for university and research. the data of sites s1, s2 and s3 were produced in the framework of prisma-1 “programma di ricerca e sperimentazione per la salvaguardia del mare adriatico” project, financially supported by the italian ministry for university and research. the data for the site mi were produced in the framework of prisma-1 and mat “programma di monitoraggio e studio sui processi di formazione delle mucillagini nell’adriatico e nel tirreno” project, financially supported by the italian ministry for the environment and the safeguard of the territory and the sea. the data for the site cs were produced in the framework of “intervento 72 – campo sperimentale in no nco mm er cia l u se on ly m. giani et al.56 mare” project, financially supported by the veneto region and coordinated by arpavregional agency for prevention, environment and energy of veneto (italy). we thank all the technicians and researchers of cnr, of icram and of the other institutions who collaborated to the field and laboratory work and contributed to produce the data that were used in the present work. references ambrose rf, anderson tw, 1990. influence of an artificial reef on the surrounding infaunal community. mar. biol. 107: 41-52. arpav, 2018. [stime delle portate alla foce dei principali corsi d’acqua del veneto per il triennio 2014-16, ai fini della valutazione dei carichi veicolati in adriatico].[report in italian]. nota tecnica n. 06/18: 39 pp. autorità di bacino dell’adige e dei fiumi dell’alto adriatico, 2009. [documento di piano, pressioni bacino del fiume sile, pp. 1-25].[report in italian]. autorità di bacino dell’adige e dei fiumi dell’alto adriatico, venezia. autorità di bacino dell’adige e dei fiumi dell’alto adriatico, 2009. [documento di piano, pressioni bacino del fiume piave, pp. 1-66].[report in italian]. autorità di bacino dell’adige e dei fiumi dell’alto adriatico, venezia. bernardi f, berton a, bastianini m, socal g, acri f, 2004. phytoplankton succession in a coastal area of the nw adriatic over a 10-years sampling period (1990-1999). cont. shelf res. 24:97-115. berger wh, wefer g, 1990. export production: seasonality and intermittency, and paleoceanographic implication. palaeogeogr. palaeoclimatol. palaeoecol. 89:245-254. bertuzzi a, faganeli j, welker c, brambati a, 1997. benthic fluxes of dissolved inorganic carbon, nutrients and oxygen in the gulf of trieste (northern adriatic). water air soil poll. 99:305-314. boldrin a, miserocchi s, rabitti s, turchetto m, balboni v, socal g, 2002. particulate matter in the southern adriatic and ionian sea: characterization and downward fluxes. j. marine syst. 33-34:389-410. boldrin a, carniel s, giani m, marini m, bernardi aubry f, campanelli a, grilli f, russo a, 2009. effects of bora wind on physical and biogeochemical properties of stratified waters in the northern adriatic. j. geophys. res. 114: c08s92. doi: 10.1029/2008jc004837. boldrin a, langone l, miserocchi s, turchetto m, acri f, 2005. po river plume on the adriatic continental shelf: observations on dispersion and sedimentation dynamics of dissolved and suspended matter during different river discharge rates. mar. geol. 222-223:135-158. boldrin a, de lazzari a, pugnetti a, bazzoni am, cassin d, turchetto m, 2006. [produttività primaria, flussi verticali di materiale particellato ed export di carbonio organico nell’area del campo sperimentale, p. 70-82]. in: [campo sperimentale in mare: prime esperienze nel veneto relative a elevazioni del fondale con materiale inerte].[book in italian]. arpav-regione del veneto. bogunović b, malačič v, 2009. circulation in the gulf of trieste: measurements and model results. il nuovo cimento 31:301-326. brambati a, ciabatti m, fanzutti gp, marabini f, marocco r, 1983. a new sedimentological textural map of the northern and central adriatic sea. boll. oceanogr. teor. app. i 267-271. brambati a, bregant d, lenardon g, stolfa d, 1973. transport and sedimentation in the adriatic sea. museo friulano di storia naturale, pubbl. n. 20, udine: 60 pp. brando ve, braga f, zaggia l, giardino c, bresciani m, matta e, bellafiore d, ferrarin c, maicu f, benetazzo a, bonaldo d, falcieri fm, coluccelli a, russo a, carniel s, 2015. high-resolution satellite turbidity and sea surface temperature observations of river plume interactions during a significant flood event. ocean sci. 11:909-920. brondi a, ferretti o, anselmi b, falchi g, 1979. [analisi granulometriche e mineralogiche dei sedimenti fluviali e costieri del territorio italiano].[article in italian]. boll. soc. geol. ital. 98:293-326. buesseler ko, antia an, chen m, fowler sw, gardner wd, gustafsson o, harada k, michaels af, rutgers van der loeff m, sarin m, steinberg dk, trull tw, 2007. an assessment of the use of sediment traps for estimating upper ocean particle fluxes. j. mar. res. 65:345-416. colella s, falcini f, rinaldi e, sammartino m, santoleri r, 2016. mediterranean ocean colour chlorophyll trends. plos one 11:e0155756. doi:10.1371/journal.pone.0155756. correggiari a, trincardi f, langone l, roveri m, 2001. styles of failure in late holocene highstand prodelta wedges on the adriatic shelf. j. sediment. res. part b 71:218-236. cozzi s, falconi cf, comici c, čermelj b, kovac n, turk v, giani m, 2012. recent evolution of river discharges in the gulf of trieste and their potential responses to anthropogenic pressure and climate change. estuar. coast. shelf sci. 115:14-24. cozzi s, giani m, 2011. river water and nutrient discharges in the northern adriatic sea: current importance and long-term changes. cont. shelf res. 31:1881-1893. danovaro r, gambi c, manini e, fabiano m, 2000. meiofauna response to a dynamic river plume front. mar. biol. 137: 359-370. de lazzari a, boldrin a, cassin d, 2006. [caratteristiche tessiturali e chimiche dei sedimenti superficiali nell’area a barriere artificiali del “campo sperimentale” p. 85-96]. in: [campo sperimentale in mare: prime esperienze nel veneto relative a elevazioni del fondale con materiale inerte].[book in italian]. arpav-regione del veneto. d’ortenzio f, ribera d’alcala m, 2009. on the trophic regimes of the mediterranean sea: a satellite analysis. biogeosciences 6:139-148. falcão m, santos mn, drago t, serpa d, monteiro c, 2009. effect of artificial reefs (southern portugal) on sedimentwater transport of nutrients: importance of the hydrodynamic regime. estuar. coast. shelf sci. 83:451-459. faganeli j, 1989. sedimentation of particulate nitrogen and amino acids in shallow coastal waters (gulf of trieste, northern adriatic). mar. chem. 26:67-80. faganeli j, malej a, pezdic j, malacic v, 1988. c: n: p ratios and stable c isotopic ratios as indicators of sources of organic matter in the gulf of trieste (northern adriatic). oceanol. acta 11:377-382. no nco mm er cia l u se on ly fluxes of particulate matter, carbonates, organic carbon and nitrogen in the northern adriatic continental shelf 57 faganeli j, planinc r, pezdic j, smodis b, stegnar p, ogorelec b, 1991. marine geology of the gulf of trieste (northern adriatic): geochemical aspects. mar. geol. 99:93-108. fain amv, ogston as, sternberg rw, 2007. sediment transport event analysis on the western adriatic continental shelf. cont. shelf res. 27:431-451. fonda umani s, del negro p, larato c, de vittor c, cabrini m, celio m, falconi c, tamberlich f, azam f, 2007. major inter-annual variations in microbial dynamics in the gulf of trieste (northern adriatic sea) and their ecosystem implications. aquat. microb. ecol. 46:163-175. fox jm, hill ps, milligan tg, ogston as, boldrin a, 2004. floc fraction in the waters of the po river prodelta. cont. shelf res. 24:1699-1715. franco p, jeftic l, malanotte rizzoli p, michelato a, orlic m, 1982. descriptive model of the northern adriatic. oceanol. acta 5: 379-389. frignani m, langone l, ravaioli m, sorgente d, alvisi f, albertazzi s, 2005. fine-sediment mass balance in the western adriatic continental shelf over a century time scale. mar. geol. 222-223:13-133. gasith a, 1975. tripton sedimentation in eutrophic lake–simple correction for suspended matter. verhand.nintern. verein. limnol. 19:116-122. giani m, alberighi l, bacci c, boldrin a, puddu a, rabitti s, 1999. a preliminary approach to the carbon cycling in the northern adriatic sea, p. 251-267. in: t.s. hopkins, a. artegiani, g. cauwet, d. degobbis and a. malej (eds.), ecosystem research report n. 32: the adriatic sea. european commission, brussel. giani m, berto d, rampazzo f, savelli f, alvisi f, giordano p, ravaioli m, frascari f, 2009. origin of sedimentary organic matter in the north-western adriatic sea. estuar. coast. shelf sci. 84:573-583. giani m, boldrin a, matteucci g, frascari f, gismondi m, rabitti s, 2001. downward fluxes of particulate carbon, nitrogen and phosphorus in the north western adriatic sea. sci. tot. environ. 266:125-134. giani m, bortolozzo f, gismondi m, manzueto l, righi s, romano e, savelli f, trentin a, 1997. [carbonio azoto e fosforo nei sedimenti dell’adriatico: distribuzione e variazioni dei rapporti c:n:p nei processi sedimentari recenti]. rapporto finale. programma di ricerca e sperimentazione per la salvaguardia del mare adriatico (prisma), sotto-progetto “valutazione dei carichi esogeni – valutazione dei flussi da e verso i fondali”, icram, chioggia, pp. 1-80. [in italian]. giani m, djakovac t, degobbis d, cozzi s, solidoro c, fonda umani s, 2012. recent changes in the marine ecosystems of the northern adriatic sea. estuar. coast. shelf sci. 115:1-13. giani m, rampazzo f, berto d, 2010. humic acids contribution to sedimentary organic matter on a shallow continental shelf (northern adriatic sea). estuar. coast. shelf sci. 90:103-110. giani m, savelli f, berto d, zangrando v, ćosović b, vojvodić v, 2005. temporal dynamic of dissolved and particulate organic carbon in the northern adriatic sea in relation to the mucilage events. sci. tot. environ. 353:126-138. giani m, savelli f, boldrin a, 2003a. temporal variability of particulate organic carbon, nitrogen and phosphorus in the northern adriatic sea. hydrobiologia 494: 319-325. giani m, savelli f, ciriaco s, odorico r, 2003b. [flussi verticali di particellato sedimentabile nel golfo di trieste], p. 367386. in: [programma di monitoraggio e studio sui processi di formazione delle mucillagini nell’adriatico e nel tirreno. rapporto finale].[report in italian]. mat, icram, roma. gilmartin m, revelante n, 1983. the phytoplankton of the adriatic sea: standing crop and primary production. thalassia jugoslavica 19:173-188. giordani p, helder w, koning e, miserocchi s, danovaro r, malaguti a, 2002. gradients of benthic-pelagic coupling and carbon budgets in the adriatic and northern ionian sea. j. marine syst. 33-4: 365-387. godrijan j, young jr, marić pfannkuchen d, precali r, pfannkuchen m, 2018. coastal zones as important habitats of coccolithophores: a study of species diversity, succession, and life-cycle phases. limnol. oceanogr. 63:1692-1710. goñi ma, hedges ji, 1995. sources and reactivities of marine derived organic matter in coastal sediments as determined by alkaline cuo oxidation. geochim. cosmochim. acta 59:2965-2981. goñi ma, thomas ka, 2000. sources and transformations of organic matter in surface soils and sediments from a tidal estuary, north inlet, south carolina, usa. estuaries 23: 548-564. hedges ji, stern jh, 1984. carbon and nitrogen determinations of carbonate-containing solids. limnol. oceanogr. 29:657-663. heussner s, ratti c, carbonne j, 1990. the pps3 time-series sediment trap and the trap sample processing techniques used during the ecomarge experiment. cont. shelf res. 10:943-958. lampitt rs, antia an, 1997. particle flux in deep seas: regional characteristics and temporal variability. deep-sea res. pt. i 44:1377-1403. lazzari p, solidoro c, ibello v, salon s, teruzzi a, beranger k, colella s, crise a, 2012. seasonal and inter-annual variability of plankton chlorophyll and primary production in the mediterranean sea: a modelling approach. biogeosciences 9:217-233. lipizer m, partescano e, rabitti a, giorgetti a, crise a, 2014. qualified temperature, salinity and dissolved oxygen climatologies in a changing adriatic sea. ocean sci. 10: 771-797. langone l, miserocchi s, boldrin a, bonaldo d, carniel s, chiggiato j, turchetto m, borghini m, tesi t, 2016. dynamics of particles along the western margin of the southern adriatic: processes involved in transferring particulate matter to the deep basin. mar. geol. 375: 28-43. luchetta a, cantoni c, catalano g, 2010. new observations of co2-induced acidification in the northern adriatic sea over the last quarter century. chem. ecol. 26:1-17. ludwig w, dumont e, meybeck m, heussner s, 2009. river discharges of water and nutrients to the mediterranean and black sea: major drivers for ecosystem changes during past and future decades? prog. oceanogr. 80:199-217. malačič v, petelin b, 2009. climatic circulation in the gulf of trieste (northern adriatic). j. geophys. res. 114:c07002. doi: 10.1029/2008jc004904. marty jc, goutx m, guigue c, leblond n, raimbault p, 2009. short-term changes in particulate fluxes measured by drifting sediment traps during end summer oligotrophic no nco mm er cia l u se on ly m. giani et al.58 regime in the nw mediterranean sea. biogeosciences 6:887-899. matteucci g, frascari f, 1997. fluxes of suspended materials in the north adriatic sea (po prodelta area). water air soil poll. 99:557-572. mauri e, poulain p-m, juznic-zonta z, 2007. modis chlorophyll variability in the northern adriatic sea and relationship with forcing parameters. j. geophys. res. 112:c03s11. doi:10.1029/2006jc003545. miquel jc, fowler s, hamilton t, heilrnann j, la rosa j, carrol m, 1999. carbon fluxes and export in the northern and central adriatic sea measured with drifting sediment traps, p. 509-524. in: t.s. hopkins, a. artegiani, g. cauwet, d. degobbis and a. malej (eds.), ecosystem research report n. 32: the adriatic sea. european commission, brussel. miquel jc, fowler sw, la rosa j, buat-menard p, 1994. dynamics of the downward flux of particles and carbon in the open northwestern mediterranean sea. deep-sea res. pt. i 41:243-261. miquel j-c, martín j, gasser b, rodriguez-y-baena a, toubal t, fowler sw, 2011. dynamics of particle flux and carbon export in the northwestern mediterranean sea: a two decade time-series study at the dyfamed site. prog. oceanogr. 91:461-481. moodley l, heip chr, middelburg jj, 1998. benthic activity in sediments of the northwestern adriatic sea: sediment oxygen consumption, macroand meiofauna dynamics. j. sea res. 40:263-280. mozetič p, solidoro c, cossarini g, socal g, precali r, francé j, bianchi f, de vittor c, smodlaka n, umani sf, 2010. recent trends towards oligotrophication of the northern adriatic: evidence from chlorophyll a time series. estuar. coast. 33:362-375. ogorelec b, mišič m, faganeli j, 1991. marine geology of the gulf of trieste (northern adriatic): sedimentological aspects. mar. geol. 99:79-92. ogrinc n, faganeli j, pezdic j, 2003. determination of organic carbon remineralization in near-shore marine sediments (gulf of trieste, northern adriatic) using stable carbon isotopes. org. geochem. 34:681-692. perdue em, koprivnjak jf, 2007. using the c/n ratio to estimate terrigenous inputs of organic matter to aquatic environments. estuar. coast. shelf sci. 73:65-72. pettine m, patrolecco l, camusso m, crescenzio s, 1998. transport of carbon and nitrogen to the northern adriatic sea by the po river. estuar. coast. shelf sci. 46:127-142. puddu a, la ferla r, allegra a, bacci c, lopez m, oliva f, pierotti c, 1998. seasonal and spatial distribution of bacterial production and biomass along a salinity gradient (northern adriatic sea). hydrobiologia 363:271-282. pugnetti a, camatti e, mangoni o, morabito g, oggioni a, saggiomo a, 2006. phytoplankton production in italian freshwater and marine ecosystems: state of the art and perspectives. chem. ecol. 22:s49-s69. puškaric s, fowler sw, miquel jc, 1992. temporal changes in particulate flux in the northern adriatic sea. estuar. coast. shelf sci. 35:267-287. ramondenc s, goutx m, lombard f, santinelli c, stemmann l, gorsky g, guidi l, 2016. an initial carbon export assessment in the mediterranean sea based on drifting sediment traps and the underwater vision profiler data sets. deep-sea res. pt. i 117:107-119. ravaioli m, alvisi f, menegazzo vitturi l, 2003. dolomite as a tracer for sediment transport and deposition on the northwestern adriatic continental shelf (adriatic sea, italy)., cont. shelf res. 23:1359-1377. rizzo g, boldrin a, de lazzari a, bressan m, 2009. swimmers in trappole per il sedimento nel nord adriatico swimmers in sediment traps in the northern adriatic sea. biol. mar. mediterr. 16:154-157. smodlaka n, 1986. primary production of the organic matter as an indicator of the eutrophication in the northern adriatic sea. sci. total environ. 56:211-220. solidoro c, bastianini m, bandelj v, codermatz r, cossarini g, melaku canu d, ravagnan e, salon s, trevisani s, 2009. current state, scales of variability, and trends of biogeochemical properties in the northern adriatic sea. j. geophys. res. 114:c07s91. doi: 10.1029/2008jc004838. stavrakakis s, chronis g, tselepides a, heussner s, monaco a, abassi a, 2000. downward fluxes of settling particles in the deep cretan sea (ne mediterranean). prog. oceanogr. 46: 217-240. tesi t, langone l, ravaioli m, giglio f, capotondi l, 2012. particulate export and lateral advection in the antarctic polar front (southern pacific ocean): one-year mooring deployment. j. marine syst. 105-108:70-81. tesi t, miserocchi s, goñi ma, langone l, boldrin a, turchetto m, 2007. organic matter origin and distribution in suspended particulate materials and surficial sediments, from the western adriatic sea (italy). estuar. coast. shelf sci. 73:431-446. turchetto m, boldrin a, rabitti s, acri f, comaschi a, 2002. particulate matter and downward flux in the northern and central adriatic sea. chem. ecol. 18:127-141. usgofs (global ocean flux study), 1989. sediment trap technology and sampling. report of the u.s. gofs working group on sediment trap technology and sampling. whoi: 94 pp. viličić d, djakovac t, burić z, bosak s, 2009. composition and annual cycle of phytoplankton assemblages in the northeastern adriatic sea. bot. mar. 52:291-305. wang xh, pinardi n, 2002. modeling the dynamics of sediment transport and resuspension in the northern adriatic sea. j. geophys. res. 107:3225. doi: 10.1029/2001jc001303. wang xh, pinardi n, malicic v, 2007. sediment transport and resuspension due to combined motion of waves and currents in the northern adriatic sea during a bora event in january 2001: a numerical modelling study. cont. shelf res. 27:613-633. wassman p, fonda umani s, ypma je, reigstad m, cok s, salvi c, cauwet g, 1999. suspended biomass and vertical flux in the gulf of trieste (northern adriatic sea): two summer scenarios, p. 537-55. in: t.s. hopkins, a. artegiani, g. cauwet, d. degobbis and a. malej (eds.), ecosystem research report n. 32: the adriatic sea. european commission, brussel. zennaro e, scatto d, patti s, 2006. [prima realizzazione in veneto di elevazioni del fondale con materiale inerte: il “campo sperimentale in mare”], p. 35-45. [campo sperimentale in mare: prime esperienze nel veneto relative a elevazioni del fondale con materiale inerte].[book in italian]. arpavregione del veneto. no nco mm er cia l u se on ly layout 1 introduction cyanobacteria are a common and naturally occurring component of freshwater environments, although they can be found in all terrestrial and aquatic ecosystems (whitton, 2012). they are important primary producers and play a key role in ecosystem functioning and biodiversity. however, they can pose risks to environment and aquatic consumers through the production of cyanotoxins (ctx), a large group of toxins that comprise microcystins (mcs), which are the most numerous and studied chemical variants (buratti et al., 2017). during dense blooms, which are increasingly occurring due to eutrophication and climate changes (paerl and paul, 2012; planas and paquet, 2016), ctx can reach very high concentrations, affecting both humans and aquatic animals, as well as wild and livestock animals (funari and testai, 2008; moreira et al., 2013; hilborn and beasley, 2015; testai et al., 2016b; wood, 2016; buratti et al., 2017). reports of animal poisonings attributable to ctx have been documented worldwide for more than a century. a diverse range of animals has been affected from dogs, cattle and fish, to flamingos, bats and bees (carbis et al., 1995; frazier et al., 1998; briand et al., 2003; stewart et al., 2008). some animals appear to be attracted by cyanobacteria in water and dried crusts on top of the water, even when clean water was plainly accessible (codd et al., 1992; lopez rodas and costas, 1999). oral human exposure to ctx can occur through the ingestion of drinking water from a contaminated source or through the ingestion of inadveradvances in oceanography and limnology, 2017; 8(1): 71-86 article doi: 10.4081/aiol.2017.6352 this work is licensed under a creative commons attribution-noncommercial 4.0 international license (cc by-nc 4.0). cyanobacterial dynamics and toxins concentrations in lake alto flumendosa, sardinia, italy mara stefanelli,1,2 simona scardala,2 piera angela cabras,3 andrea orrù,3 susanna vichi,2 emanuela testai,2 enzo funari,2 maura manganelli2* 1inail (national institute for insurance against accidents at work), via roberto ferruzzi 38/40, 00143 rome; 2istituto superiore di sanità (national health institute), department of environment and health, viale regina elena 299, 00161 rome; 3istituto zooprofilattico sperimentale della sardegna (experimental zootechnic institute of sardinia), via fratelli kennedy 2, 08100 nuoro, italy *corresponding author: maura.manganelli@iss.it abstract seasonal blooms of cyanobacteria (cb) are a typical feature of lake alto flumendosa (sardinia, italy). the waters of this lake are used for drinking water supply, for agricultural and industrial uses, and fish farming activities. since cyanotoxins are not monitored in edible organisms, diet could be a relevant route of human exposure. cb also represent a threat for the health of wild and domestic animals that use lake water for beverage. therefore, to characterize the cb community and assess the risk for human and animal population, cb dynamic, mcyb+ fraction, and microcystins (mcs) concentration have been followed monthly for 18 months, in three stations. results confirmed the presence of several toxigenic species. planktothrix rubescens dominated between august 2011 and april 2012 (3.5×106 cells l–1), alternating with woronichinia naegeliana (8×106 cells l–1) and microcystis botrys (9×105 cells l–1). dolichospermum planctonicum was always present at low densities (104 cells l–1). mcs were detected, at values well below the 1 µg l–1 threshold of who for drinking water. the molecular analysis of mcyb gene for p. rubescens indicated the presence of a persistent toxic population (average 0.45 mcyb/16s rdna). highly significant linear regressions were found between p. rubescens and the sum of the demethylated mc variants, and between m. botrys and the sum of mc-lr and mc-la, also when co-occurring, suggesting that these two species were responsible for different mc patterns production. the regression lines indicated a quite stable mc cell quota. however, in some spotted samples very different values were obtained for both mc concentrations and cell quota (from 10-fold lower to 30-40-fold higher than the ‘average’) showing an unexpected significant variability in the rate of toxin production. the relatively low cell densities during the monitoring period is consistent with the low-to absent mc contamination level found in trout muscle; however, the analytical method was affected by low recovery, probably due to mc-protein binding. our results show that, during the study period, no risk of exposure for the human and animal population occurred. however, the persistence of a complex cb community characterised by a significant toxic fraction suggests the need for periodic monitoring activity. particularly, the hidden deep summer p. rubescens blooms, located where water is taken for drinking water supply, and m. botrys, able to produce the most toxic mc variants with high cell quota, should be kept under control. the documentation and interpretation of sudden changes in toxins concentrations deserve special attention. this is particularly relevant in proximity of fish farming plants and water catchment sites. key words: toxic cyanobacteria; cyanotoxins; microcystins; human and animal exposure; health risk. received: 19 october 2016. accepted: 3 march 2017.no nco mm er cia l u se on ly m. stefanelli et al.72 tently swallowed contaminated water during recreational activities (funari and testai, 2008; testai et al., 2016a). furthermore, humans can be orally exposed through the ingestion of cyanobacteria-based food supplements (saker et al., 2005; vichi et al., 2012) or food items such as fish and shellfish with bioaccumulated toxins (e.g., through filtration of contaminated water) (ibelings and chorus, 2007; berry, 2013). additional routes of exposure are dermal contact and accidental inhalation during recreational activities in waters subjected to a toxic bloom. in sardinia, more than 90% of the drinking water originates from artificial lakes which are generally eutrophic (lugliè et al., 2013), the ideal environment for cyanobacterial blooms. indeed, cyanobacteria have become dominant in many reservoirs over time (sechi and lugliè, 1992, 1996; aktan et al., 2009; pulina et al., 2011). however, only a few studies have assessed the presence of ctx, namely mcs (messineo et al., 2009; sulis et al., 2014; mariani et al., 2015). furthermore, a toxic strain of microcystis aeruginosa (kützing) kützing, associated with a fish kill event, was isolated in lake liscia (pellegrini et al., 1995), and a toxic bloom of planktothrix rubescens (formerly oscillatoria rubescens) (de candolle ex gomont) anagnostidis and komárek was documented in 1986 in the flumendosa reservoir (loizzo et al., 1988). in the last 30 years, the flumendosa reservoir has been used as the main water supply for civilian, agricultural and industrial uses in southern sardinia (botti et al., 2001). in addition, a trout farming plant is located nearby, whereas water is utilised also by wild animal and livestock for beverage. in the 1990s, cyanophyceae, together with bacillariophyceae and chlorophyceae, were the dominant classes in the flumendosa reservoir system (lugliè et al., 1997), with cb representing 90-100% of phytoplankton since 2002 (sulis et al., 2014). the cb with the highest density described so far are the planktothrix and microcystis genera, which have been responsible for the most frequent onset of blooms, while the dolichospermum genus, except for a bloom in 2003, appeared sporadically (sulis et al., 2014). to our knowledge, data on mcs in this lake have been assessed only in two occasions and were associated with the presence of p. rubescens (messineo et al., 2009; sulis et al., 2014). a recent 18-month study on four reservoirs in northern sardinia showed significantly different seasonal variations in cb community composition, with several toxigenic species, and large variability in mcs concentrations (mariani et al., 2015). the results highlighted the need to increase the knowledge of these aspects in each reservoir, whose use could affect human and animal health. in june 2010, a mass mortality of fish occurred in the flumendosa reservoir, concurrently with an extensive surface bloom of p. rubescens. the cause for fish mortality was potentially attributed to a combination of oxygen deficiency and cb toxins. following this event, our study was aimed: i) to characterize the cb community of the reservoir; ii) to verify the applicability of a linear model used in single species blooms to predict their toxicity (salmaso et al., 2014); and iii) to assess the potential risk for humans and animals. we followed the dynamics of cyanobacteria for 18 months and determined mcs concentrations, toxicity profile (mc variants concentration), frequency of toxic genotype and accumulation of mcs in farmed trout. methods description of the site and sampling strategy the reservoir of the alto flumendosa, located in centre-eastern sardinia (italy) (latitude: 39°42’38’’ n; longitude: 9°17’25’’ e) (fig. 1), is an artificial lake (volume 64×106 m3) with a surface area of 9 km2 (begliutti et al., 2007) and a maximal depth of around 50 m. it is a complex and multiuse water system, which interconnects with other reservoirs (flumendosa-campidano). the reservoirs are linked in a cascading sequence and from each one pressure pipelines and open channels guarantee water for residential use and irrigation of the campidano plain (sechi and sulis, 2009). once used for the production of electrical energy in three underground stations, the water is conveyed through a system of tunnels in the lake teaula, about 286 m asl, east of villagrande. at this point, waters are directed through the tortolì flat for irrigation and for civilian uses (e.g., drinking water) (cabras et al., 2013). water samples were collected in three sites: (1) zattere (maximum depth 50 m) at about 10 m from a trout’s floating cages farming plant; (2) middle site in the centre of the lake (maximum depth 25 m), in front of the water offtake site (collecting water also for drinking purposes); and (3) rio osiana (maximum depth 7 m) close to the inflow of a small temporary tributary stream (fig. 1). during our sampling period (october 2011 may 2013), the stream periodically underwent dry phases. after oct 2012, the site 3 was no longer sampled, since the values of the studied parameters were not different from the other sites. since in some periods some cb species density was very low, two different strategies of sampling were used to monitor their temporal dynamics: i) discrete samples were collected monthly in the whole sampling period (18 months) in the three stations at the surface (s1, s2, s3) and in the stations 1 and 2 at 20 m depths (p1, p2; depth of the water offtake site, which is important to check when p. rubescens occurs) (manganelli et al., 2016) using a van dorn bottle; ii) periodically, integrated samples were collected with a 20 µm phytoplankton net on a water column of 30 m (r1), 25 m (r2) and 7 m (r3). samples were collected and stored in acid clean polycarbonate bottles, in cold and dark conditions until the arrival in the laboratory where they were aliquoted for the different analysis. no nco mm er cia l u se on ly cyanotoxins in a sardinian reservoir 73 five fish samples were collected from november 2011 to september 2012, every two-three months, from the trout farming plant (2-3 animals for sampling). physical measurements and chlorophyll-a analysis chlorophyll-a (chl-a) was extracted according to jespersen and christoffersen (1987). samples were filtered onto gf/f filters (whatman) and stored in 96% ethanol in the dark at room temperature over-night. the fluorescence of the extract was measured with a turner 10au-005 fluorimeter (holm-hansen et al., 1965). temperature was read from a thermometer inside the sampling bottle, as soon as the sample arrived at the surface, and ph was measured in the field with a ph meter probe. cyanobacteria abundance and isolation cyanobacteria taxonomic determination was carried out according to komárek and anagnostidis (1999, 2005) and komárek and zapomělová (2007). cyanobacteria abundance was determined by epifluorescence microscopy, since phycoerythrin and phycocyanin contained in cb, when examined under blue light excitation, fluoresce orange and red, respectively (walsby and avery, 1996; ernst et al., 2006). this technique, based on sample concentration on filters, allows the concentration of large volumes in few minutes and the counts of species present within a large range of densities (hallegraeff et al., 2004). fifty ml of formaldehyde-fixed samples (final concentration 4%) were filtered onto 5 µm black membrane filters (25 mm diameter) (whatman). the number of filaments and colonies was counted by autofluorescence under blue light with an upright microscope (olympus bx51) at 100x. two replicate filters per sample were counted, with a variation between filters generally well within 20%. the number of cells per filament/colony was determined by average on 50 filaments/colony on both filters, using an image analysis software (imagej, http://rsbweb.nih.gov/ij/). monoclonal strains of p. rubescens were isolated as follows. the filaments were plated on bg-11 medium 1% agar (difco) plates. the plates were then incubated at 14°c, at a photosynthetic photon flux density of 10 µmol m–2 s–1 in a light:dark cycle (16:8h). filaments were isolated by micromanipulation according to rippka (1988) fig. 1. the flumendosa reservoir and sampling sites. no nco mm er cia l u se on ly m. stefanelli et al.74 and incubated at the same conditions described above, in liquid bg-11 medium. we also tried to isolate colonial cyanobacteria, like woronichinia naegeliana (unger) elenkin, as described by sedmak et al. (2008). the samples were concentrated in test tubes under natural light. colonies floated towards the surface due to their buoyancy mechanism. the remaining phytoplankton together with zooplankton sank to the bottom. the surface layer with cyanobacterial material was collected and re-examined by microscopy resulting monospecific (over 99%). the colonies were then incubated in liquid bg-11 at the same conditions as above. however, all isolation attempts failed. w. naegeliana did not persist in culture more than one month and no further analysis was carried out. phytoplankton abundance samples for the identification and counting of phytoplankton other than cyanobacteria were fixed with formaldehyde (final concentration 4%) and/or lugol’s solution (final concentration 1%). ten ml samples were sedimented according to the utermhöl method (1958) and counted under an inverted microscope (olympus ix50) at 40x and 400x magnification, in bright-dark field. detection of microcystins elisa total (intraand extra-cellular content) mcs were analysed in untreated water samples by an enzyme-linked immunosorbent assay (elisa). after three freeze/thaw cycles, the samples were analysed using an elisa kit (microcystins adda-elisa microtiter plate, envirologix, portland, me, usa) according to the manufacturer’s protocol. microscope observation of freeze thawed samples confirmed the complete lysis of cells. the results were expressed as mc-lr equivalents (mc-eq). methanol extracts from fish samples were also tested, after being dried and re-dissolved in water, with minor modification from papadimitriou et al. (2012). briefly, 5 g of fish muscle from each animal were added with 10 ml methanol, homogenized, centrifuged at 4000 rpm for 10 min and the supernatant collected. the extraction step was repeated twice and the supernatant fractions were pooled. three aliquots of 5 ml each were gently evaporated to dryness at 45°c in rotavapor. the residue was reconstituted in 500 µl of distilled water. samples were considered positive when the mc concentration was higher than the detection limit of the method (0.10 μg l–1). lc-ms/ms analysis to discriminate the different mc variants, which have a different toxicity (funari and testai, 2008; buratti et al., 2017), samples have also been analysed by liquid chromatography tandem mass spectrometry (lc-ms/ms). the extraction procedure was different for water samples and complex matrix as tissues. water samples each sample was divided into two 500 ml aliquots: one was used to determine dissolved mcs and one to determine total mcs. for dissolved fraction detection, samples were filtered onto gf/c discs (whatman) in order to eliminate cells. the filtrate, acidified, was cleaned-up by solid-phase extraction (spe), as previously described (buratti et al., 2011). briefly, samples were purified and concentrated through ods spe cartridge, rinsed with 20% meoh (4 ml), then eluted with 4 ml meoh. the eluate was dried under a gentle nitrogen stream, and re-dissolved in 0.5 ml of acetonitrile:water (30:70) containing 0.1% (v/v) formic acid. to determine total mcs (dissolved plus intracellular mcs), the aliquot was freeze-thawed three times, to cause cell lysis with the consequent release of intracellular ctx and then treated as described above for the dissolved fraction. the intracellular content of mcs was calculated as the difference between total and dissolved concentration. fish fish tissues have been extracted as reported above for elisa assay. separation, identification and quantification of the mc variants (mc-lr, -la, -yr, -rr, -lf, -lw, -demlr, -dem-rr, for which standards were available) were performed with two micropump and an autosampler pe series 200 (perkin elmer inc., waltham, ma, usa), coupled to a triple-quadrupole mass spectrometer equipped with a turboionspray source (mds sciex, concord, canada). ms tuning and optimization were achieved by infusing mcs standards (alexis, san diego, ca, usa); detection was carried out using multiple reaction monitoring (mrm) mode. mcs separation, in mrm analysis, was conducted as previously described (buratti et al., 2011). all samples were run in mrm mode, utilizing the specific fragmentation reaction of single or doubly protonated mcs to m/z 135, the specific fragment associated to the adda fragment. the amount of mcs was quantified by the external standard quantification procedure, referring to a calibration straight line with 4–6 known amounts of the analytical standards in solvent (coefficient of determination, r2≥0.98), for water samples, and in the blank fish matrix, for fish samples, in order to exclude any interference due to matrix effect. indeed, salts that are hardly removed during extraction procedures and sample preparation may influence samples ionization, causing a quenching of the signal for some congeners whereas for others the signal can be overcompensated (vichi et al., 2012). in fish samples, recovery was also investigated no nco mm er cia l u se on ly cyanotoxins in a sardinian reservoir 75 spiking preand post-extraction with known amounts of standard mcs. we found around 80% recovery for the different mc variants tested, except for mc-rr, mc-demrr and mc-la, for which recovery was never higher than 50%. for the applied method, the loq ranged between 2 ng l–1 (mc-rr) and 9 ng l–1 (mc-lw) for water samples, and between 5 ng g–1 (mc-lr) and 15 ng g–1 (mc-lf) for fish samples. molecular investigation each field sample (250-500 ml) was filtered on 0.2 µm supor membrane filters to harvest the cells; filters were stored at -20°c. total genomic dna was extracted using the commercial dneasy plant mini kit (qiagen), following the manufacturer’s instructions. genomic dna concentration was measured by a biophotometer (eppendorf), and purity was assessed by calculating the ratio of the absorbance at 260 nm and 280 nm. molecular analyses were performed exclusively on 83 samples, where the p. rubescens cells abundance, estimated by counting before storage and dna extraction, exceeded or was equal to the threshold of 104 cells l–1, in order to obtain reliable and consistent results. a preliminary qualitative pcr was carried out to identify the mc-producing genera planktothrix and microcystis, following the methods of rantala et al. (2006) and vaitomaa et al. (2003), based on the detection of the genus-specific mc synthetase gene e (mcye). the quantitative estimation of planktothrix cells and the fraction of individuals carrying mcs biosynthesis genes were further investigated by qpcr employing a stepone™ realtime pcr system (applied biosystems), according to ostermaier and kurmayer (2009). two independent taq nuclease assays (tnas) were used, one to quantify the general population of planktothrix using the 16s rdna region as specific genus marker, and the other one to quantify the potential mc-producing genotypes within the community by checking the amplification of the first adenylation domain of mcyb (mcyba1). genotypes of the unknown samples were quantified relating their threshold cycle (ct) to the starting cell concentration, by applying the standard curve methodology. for both 16s rdna and mcyb genes the standard curves were established using serial dilutions of dna extracted from known cell concentrations of p. rubescens strains ccap1460/3 and ccap1460/10. the serial dilutions and the samples were amplified simultaneously to determine their ct values under the same experimental conditions. each measurement was performed in triplicate. a similar qpcr approach was applied to quantify the toxic fraction of microcystis (vichi et al., 2012); however, some cross reactivity was observed when primers were tested in the presence of w. naegeliana cells, which often co-occurred with microcystis in the environmental samples, thus making this method unreliable for the quantitative estimation of microcystis mcy+ cells. statistics differences between set of data were tested with anova, with the software statistica 6 (statsoft, inc. ok, usa), after testing for normality. since data from surface sampling sites were not statistically different they have been averaged and average ± sd has been considered. correlation analysis between different variables has been calculated with the software sigmaplot 9.01 (systat) on all raw data. results temperature data of temperature are shown in fig. 2. the lake was generally stratified from march-april until december. cyanobacteria abundance and isolation several potentially toxic species were detected during the monitoring period, with a variable relative abundance over the year: planktothrix rubescens, woronichinia naegeliana, microcystis botrys teiling and dolichospermum planctonicum (brunnthaler) wacklin, l. hoffmann & komárek (fig. 3). p. rubescens was always present (fig. 3a); it dominated in both surface and 20 m depth samples, from november 2011 to summer months in 2012, being significantly lower in the remaining months. during the stratification period, with surface temperatures reaching a maximum of 22.5°c, the highest densities were at 20 m depth, with values oscillating between 1.5 to a maximum of 5×106 cells l–1, according to the typical seasonal dyfig. 2. temperature variation at the surface (s, average of the 3 surface sites ± sd) and 20 m depth (p, average of the 2 depth sites ± sd). aug 2012, nov 2012, feb 2013 no sampling. no nco mm er cia l u se on ly m. stefanelli et al.76 namic of this species. after july 2012, p. rubescens density was about one order of magnitude lower than in the previous year (fig. 3a). during the stratified period, sporadic samples in s1 were collected every 5 m from surface to -20 m, confirming that p. rubescens was homogenously distributed between -10 and -20 m with higher density than surface (data not shown). w. naegeliana was numerically dominant from september to december 2012 when it reached high values on the surface (4.3-7.5×106 cells l–1) and at 20 m depth (68×106 cells l–1), that is about one order of magnitude higher than in 2011 (fig. 3b). in the remaining months, it was totally absent or present at much lower densities. m. botrys was found from the beginning of the monitoring activity until january 2012 and from may to october 2012 in surface samples and sporadically in depths samples (fig. 3c). a peak of density (up to 8×105 cells l–1, sites s1 and s2) was observed in october 2011 at the surface, when it represented the dominant species among cb (45%); it sharply decreased to less than 5% of total cb community in the following months. d. planctonicum was always present at low densities (fig. 3d). in surface samples, it never exceeded 4×104 cells l–1; densities at depths were similar or lower. the lowest values were measured in samples collected in 2013. when integrated samples were considered, m. botrys was the only species virtually absent from the lake between december 2012 and may 2013, when it was again detectable at very low density (1.7×104 cells l–1) although only in r2 (tab. 1). p. rubescens and w. naegeliana were always present at very high density (up to 1.5×109 cells l–1 and 2.4×109 cells l–1 respectively). d. planctonicum was characterized by low densities (maximum 7×106 cells l–1). phytoplankton abundance the phytoplankton community showed a high variability in both cell abundance and species composition. it consisted of green algae like chlorella sp., scenedesmus sp. and chlorococcum sp., and diatoms like aulacoseira sp., fragilaria sp. and asterionella sp. in terms of cell number, they represented on average only ~ 30% of the community (median = 15%). however, their biomass was more relevant. indeed, two significant correlations, one for summer and one for winter (r=0.73 p<0.001 and r=0.90 p<0.001, respectively), identified by two ranges of temperature, were found between the density of the whole phytoplankton community (cb and algae) and chla, used as a proxy for biomass (fig. 4). excluding noncb algae from the number of cell data, correlations were no longer significant, implying that in this lake, even if fig. 3. seasonal variations in the densities of (a) p. rubescens, (b) w. naegeliana, (c) m. botrys and (d) dolichospermum sp. at the surface (s, average of the 3 surface sites ± sd and at 20 m depth (p, average of the 2 depth sites ± sd) (y axis log scale). aug 2012, nov 2012, feb 2013 no sampling. no nco mm er cia l u se on ly cyanotoxins in a sardinian reservoir 77 cb were numerically more abundant than algae, chl-a could not be used as an index of cb abundance. detection of microcystins elisa the mcs concentrations in water were always detected in the range 0.1-1 µg l–1 (data not shown). the highest values were found in october 2011 in a surface sample, when the dominant species was m. botrys. lc-ms/ms mc-la, -lr, -dem-lr, -rr, -dem-rr were detected in all samples in different relative proportions, whereas the other tested congeners mc-lf, -ly end -lw were always below the detection limit (9 ng l–1). consistently with the elisa data, in all discrete samples mcs concentrations never exceeded 1 µg l–1 (supplementary tab. 1). the levels of mcs in integrated samples ranged between 0.4 and 100 μg l–1 (tab. 2). the highest value was measured in may 2012 when a 10-fold difference was observed between r1 and r3, although the relative proportion of the different variants was the same. the pattern of congeners in both discrete and integrated samples revealed that generally, but not always, mc-dem-rr accounted for ~92% (up to 100%) and mcdem-lr up to 11%, with the sum of the other variants <1% (tabs. 2 and 3). indeed, in some samples, the pattern of congeners was totally different, with mc-lr and mcla accounting for ~60% and 30%, respectively. the fit of linear models between single congeners and cyanobacterial species allowed distinguishing the potential producers of different mc variants. in discrete samples, highly significant correlations were observed between p. rubescens and mc-dem-rr and mc-dem-lr concentration (r2=0.81) (fig. 5a) and between m. botrys and mc-lr, -rr, -la (r2=0.86) (fig. 5b). in the integrated samples, no correlation between p. rubescens and dem-mc was observed when all data were included in the analysis. excluding the values measured in may 2012, which deviated from the main pattern, the relationship between mc concentrations and p. rubescens densities was significant (r2=0.67) (fig. 6a), with a slope not different from that obtained from the discrete samples (t-test t=0.34, p>0.5) (fig. 5a). the correlation between mc-lr and mc-la and m. botrys was still highly significant (r2=0.94) (fig. 6b). also in this case, the slopes of linear models computed on data measured in discrete and intefig. 4. variation of number of cells (phytoplankton including cyanobacteria) vs chl-a in surface sites (grey, warm season; black, cold season). tab. 1. cyanobacteria density (cells l–1×106) in integrated samples in the three sites. site date m. botrys p. rubescens w. naegeliana d. planctonicum r1 oct 2011 52.6 7.00 36.2 2.02 may 2012 nd 181 158 7.41 sept 2012 30.8 7.40 2390 0.68 oct 2012 7.04 9.87 755 0.77 dec 2012 nd 3.29 2150 2.51 jan 2013 nd 50.4 118 1.55 mar 2013 nd 94.1 13.3 0.001 apr 2013 nd 252 355 1.31 may 2013 nd 21.8 127 0.12 r2 oct 2011 71.9 13.1 100 1.63 may 2012 17.3 1510 486 5.27 apr 2013 nd 75.8 67.3 0.12 may 2013 0.02 14.8 93.3 0.07 r3 oct 2011 58.6 6.76 131 2.53 may 2012 3.52 1350 82.9 1.50 nd, not detected. no nco mm er cia l u se on ly m. stefanelli et al.78 grated samples were not statistically different (t=0.25, p>0.5). during the study, 10 strains of p. rubescens were isolated and grown in laboratory conditions. the strains were all toxic, producing only mc-dem-rr (94%±2%) and dem-lr (6%±2%). detection of mc in fish the application of elisa tests to fish muscle tissues showed a weak contamination of the examined samples (tab. 4). these results were not confirmed by the lcms/ms analyses that revealed at most not quantifiable traces of mc-lr and mc-dem-lr. molecular investigation the potential toxigenicity of the two genera microcystis and planktothrix was qualitatively confirmed by pcr in some randomly selected integrated samples with high cb cell densities. the environmental samples were amplified together with the reference strains of m. aeruginosa and p. rubescens, used as positive controls for the presence of the mcyb gene. the absence of any cross-reactivity between dna from planktothrix and microcystis and the selected primers was verified and excluded with preliminary pcr assays on each isolated environmental culture; in addition, no pcr product was obtained when the same primers were tested in samples dominated by w. naegeliana. no conclusions could be drawn on the potential toxigenicity of the w. naegeliana strains present in the lake, since no isolated reference strains (neither discriminating primers) were available. the quantitative estimate of the toxic genotypes performed on the planktothrix population by qpcr revealed a permanent and fairly constant fraction (about 40-50%) of mcyb+ cells throughout the monitoring period, with fig. 5. a) correlation between p. rubescens and mc-dem-lr + mc-dem-rr in discrete samples; the regression line is y=6.17×10–8(se 3.60×10–9) ∙ x + 8.61×10–5(se 5.65×10–3); r2=0.81, p<0.001. b) correlation between m. botrys and mclr + mc-rr + mc-la in discrete samples; the regression line is y=3.96×10–7(se 2.18×10–8) ∙ x + 3.68×10–3(se 3.38×10–3); r2=0.86, p<0.001. tab. 2. microcystins concentrations (µg l–1) in integrated samples in the three sites. site date microcystins rr dem-rr lr dem-lr la tot r1 oct 2011 nd 0.000 38.720 nd 9.180 47.900 may 2012 nd 92.097 nd 8.061 0.000 100.158 sept 2012 nd 0.865 10.381 0.266 2.573 14.085 oct 2012 0.447 nd 4.036 nd 1.345 5.829 dec 2012 nd nd 0.523 nd nd 0.523 jan 2013 nd 1.449 0.043 0.102 0.044 1.638 mar 2013 nd 4.476 nd nd nd 4.476 apr 2013 nd 6.673 0.194 1.490 nd 8.356 may 2013 nd 0.369 0.006 0.038 nd 0.413 r2 oct 2011 nd nd 39.940 nd 15.000 54.940 may 2012 nd 69.520 0.196 5.810 nd 75.526 apr 2013 nd 7.383 nd 1.372 nd 8.755 may 2013 0.012 0.692 0.015 0.057 nd 0.775 r3 oct 2011 nd nd 36.070 nd 11.880 47.950 may 2012 nd 8.338 0.221 1.181 nd 9.740 nd, not detected. no nco mm er cia l u se on ly cyanotoxins in a sardinian reservoir 79 similar proportions in the surface and deep sampling sites (48±19% vs 40±15%), and in integrated samples (56±0.18%). cell quota based on the correlations between mc congeners and cb, the toxins cell quota were calculated for p. rubescens and m. botrys. the cell quota of the dem-mc produced by p. rubescens in the discrete samples averaged 0.104±0.274 pg mc cell–1 (tab. 5). the high average and s.d. were due to two values measured in december 2012, s1 and p1, in which the cell quota was 30-40 fold higher than the rest of the values which were quite homogeneous (1.44 and 1.91 vs 0.059±0.04 pg mc cell–1 in the remaining samples). when the cell quota was calculated considering only the potentially toxic fraction (mcyb+/16s rdna), the values increased, as expected (0.130±0.076 pg mc cell–1, respectively) (tab. 5). due to a snowfall during sampling, in december 2012 no sample was collected for molecular biology analysis and therefore it was not possible to verify whether the abnormally high cell quota was related to a population 100% composed of mcyb+ cells, or to other reasons. however, even in case of a 100% toxic population the cell quota would still be ~10 times higher than usual. only assuming a possible contribution by w. naegeliana, the cell quota would be in the range of the other values, 0.010 and 0.013 pg mc cell–1; however, this possibility is unlikely, based on what is known on woronichinia toxicity. to see if the toxic fraction of p. rubescens population was a better predictor of mc concentration than the whole population, the fit of linear model between mc variants and the potentially toxic cells (mcyb+ cells) was used, but the variability exfig. 6. a) correlation between p. rubescens and mc-dem-lr + mc-dem-rr in integrated samples; the regression line is y=3.71×10–8(se 8.18×10–9) ∙ x + 0.38(se 0.68); r2=0.67, p<0.001; data from may 2012 have been excluded from the regression. b) correlation between m. botrys and mc-lr + mc-rr + mc-la in integrated samples; the regression line is y=7.94×10– 7(se 5.44×10–8) ∙ x + 1.40(se 1.58) r2=0.94 p<0.001. tab. 3. percentage of the different microcystins variants in discrete samples. site date microcystins rr dem-rr lr dem-lr la s oct 2011 0 0 60 0 40 nov 2011 4 39 44 4 9 dec 2011 1 79 9 9 1 jan 2012 0 90 0 10 0 feb 2012 0 90 0 10 0 mar 2012 0 91 0 9 0 apr 2012 0 94 0 6 0 may 2012 0 100 0 0 0 jun 2012 0 100 0 0 0 jul 2012 0 27 9 0 64 sep 2012 0 16 59 0 25 oct 2012 0 9 62 0 29 dec 2012 0 68 24 0 7 jan 2013 0 100 0 0 0 mar 2013 na na na na na apr 2013 0 100 0 0 0 may 2013 0 100 0 0 0 p oct 2011 na na na na na nov 2011 na na na na na dec 2011 2 75 12 9 2 jan 2012 0 89 0 11 0 feb 2012 0 92 0 8 0 mar 2012 0 93 0 7 0 apr 2012 0 95 0 5 0 may 2012 0 92 0 8 0 jun 2012 0 92 0 8 0 jul 2012 0 92 0 8 0 sep 2012 na na na na na oct 2012 0 100 0 0 0 dec 2012 na na na na na jan 2013 na na na na na mar 2013 0 97 3 0 0 apr 2013 0 100 0 0 0 may 2013 na na na na na s, average of the 3 surface sites; p, average of the 2 depth sites; na, not available. no nco mm er cia l u se on ly m. stefanelli et al.80 plained by the regression was much lower (54% vs 81%) (data not shown). the cell quota calculated on the integrated samples were generally consistent with the average values determined in the discrete ones, except for the values obtained in may 2012 (tab. 6), which were out of range. indeed, the cell quota from the discrete sample was 0.069 pg mc cell–1, whereas the one in r1 was almost 10 times higher (0.554 pg mc cell–1) and in r3 10 times lower ( 0.007 pg mc cell–1). the cell quota of m. botrys in discrete samples averaged 0.771±1.280 pg mc cell–1 (median 0.303 pg mc cell–1) (tab. 5); in the integrated samples cell quota averaged 0.668±0.495 pg mc cell–1, very similar to the discrete sample values (tab. 6). discussion sardinia has a typical semiarid mediterranean climate, and relies for a 57% on a complex water distribution network of surface artificial reservoirs exploited for civilian, agricultural, and industrial purposes (isri, 2006), and accounting for about 90% when only drinking water is concerned (lugliè et al., 2013). therefore a good water management plan is essential. most of the sardinian reservoirs were eutrophic (total nitrogen, 594-2057 μg l–1; total phosphorus, 23-180 μg l–1), with phytoplankton communities dominated by cb (sechi and lugliè, 1996; begliutti et al., 2007; marchetto et al., 2009). the alto flumendosa reservoir maintains an intermediate trophic level between oligotrophy and mesotrophy, as shown by the nutrient concentrations reported by cabras et al. (2013) (supplementary tab. 2), which refer to the period of the monitoring activities described in this paper. the results reported in this latter work are consistent with other previous studies (begliutti et al., 2007; sulis et al., 2014). sulis et al. (2014), who used a very large chemical and phycological dataset for the flumendosa-campidano system to calibrate a model of water quality, described a variable phytoplankton community over the years (annual averages from 1996 to 2012), with a dominance of cb (>90%) from 2002 onward. our study, which was carried out between the end of 2011 and mid 2013 fully confirmed the dominance by cb, although in terms of biomass other phytoplankton groups were also relevant. the species composition of the cb community in the flumendosa reservoir has been previously described as dynamic, with alternating dominating species (lugliè et al., 1997; meregalli et al., 2002; begliutti et al., 2007). as reported by sulis et al. (2014), p. rubescens-agardhii has been continuosly dominant (>90%) between 2002 and 2009 (peak density 35×106 cells l–1), then decreased to ~20% of the whole phytoplankton community in 2010. microcystis spp. has also been always present, generally at the onset tab. 4. elisa and lc-ms/ms results for microcystins in fish tissues. date elisa lc-ms/ms (ng g–1) (ng g–1) rr dem-rr lr dem-lr la tot 08/11/11 trout1 0.054 nd nd nd nd nd nd trout2 0.144 nd nd traces traces nd nd trout3 0.936 nd nd traces traces nd nd 28/02/12 trout1 0.200 nd nd nd nd nd nd trout2 0.132 nd nd traces traces nd nd 24/04/12 trout1 0.092 nd nd nd nd nd nd trout2 * nd nd traces traces nd nd 6/06/12 juveniles 0.048 nd nd nd nd nd nd 25/09/12 trout1 0.046 nd nd traces traces nd nd trout2 0.152 nd nd traces nd nd nd *lost sample; nd, not detected. tab. 5. average of cell quota computed for different microcystins variants (pg cell–1) in p. rubescens and m. botrys in discrete samples (see text). computations refer to the whole sampling period. species p. rubescens p. rubescens mcyb+ m. botrys discrete samples 0.104 (0.055) 0.274 0.130 (0.128) 0.076 0.771 (0.303) 1.281 n=72 n=30 n=21 mean (median) sd. no nco mm er cia l u se on ly cyanotoxins in a sardinian reservoir 81 of the blooms, while dolichospermum bloomed in 2003, but after that low densities were reported sporadically over the years (sulis et al., 2014). the overall density of cb community during our study was about 10 times lower than what was observed before (messineo et al., 2009; sulis et al., 2014), but an even more complex alternation of species was observed. p. rubescens was always present and generally dominant until summer 2012, with the typical trend of depth blooms in summer and more homogeneous distribution along the water column during the other seasons (kurmayer et al., 2016). these results confirmed the need to sample the metalimnetic layers, especially when they are the source of drinking water supply (manganelli et al., 2010, 2016). the dominating species alternating with p. rubescens was w. naegeliana, previously described in the lake only in 1995 and 1996 (meregalli et al., 2002). blooms of w. naegeliana, particularly in late summer and autumn, are not infrequent in temperate lakes, and have been reported in northern europe (including poland, the czech republic, sweden, finland, belgium, russia), north america (united states) and australia (komárek and anagnostidis, 1999; bucka and wilk-woźniak, 2002; annadottér et al., 2005; rajaniemiwacklin et al., 2005). furthermore, at the beginning of our study, m. botrys was the dominating species, then disappearing in the following months. in addition, d. planctonicum was never dominant but always present as supported by data from integrated samples on the 20 m water column, useful to detect the presence of cb species at low density. since during the winter months m. botrys was undetectable also in the integrated samples, its re-appearance during the summer months can be explained considering that colonies overwintered on the bottom sediments (reynolds et al., 1981). we tested whether cb abundance could be a good proxy for mc concentration in the lake, as it has been shown by many studies on planktothrix (briand et al., 2005; catherine et al., 2008; dolman et al., 2012; salmaso et al., 2014). a prerequisite for this relationship to be significant is the presence of a single dominant species (salmaso et al., 2014). in the present study, the toxic species occurring simultaneously were compared with the different mc-variants, analysed separately. the pattern of mc variants was demonstrated to be highly dependent on the potentially producing species: highly significant linear regressions were obtained between p. rubescens and the demethylated form of mc-rr and mclr, and between m. botrys and mc-la and mc-lr, in both discrete and integrated samples. each species was characterized by a different slope (t-test, t=2.79, p<0.01). the determination of linear regressions with the same slope in both type of samples, for each species, reinforces the validity of our results, since in the integrated samples values of cell density and mc concentration were well above the detection limits. in the elaboration of p. rubescens integrated data, results from may 2012 were not considered, being totally out of range. these results allowed to associate the potential producing species with the different mc variants and furthermore to obtain a model which explained more than 80% of the variation in mc concentration in this system. the association between p. rubescens and mc-dem-rr and -dem-lr was also confirmed by the 10 monoclonal isolated strains producing only the two variants, and is in line with data from other italian lakes (messineo et al., 2006; manganelli et al., 2010, 2016). even if no correlation between mcs and w. naegeliana was found, we cannot definitely rule out its ability to protab. 6. cell quota of microcystins variants (pg cell–1) in p. rubescens and m. botrys in integrated samples in the three sites. site date species p. rubescens p. rubescens mcyb+ m. botrys r1 oct 2011 0.000 0.000 0.911 may 2012 0.554 1.152 nd sept 2012 0.153 na 0.421 oct 2012 0.000 0.000 0.828 dec 2012 0.000 0.000 nd jan 2013 0.031 na nd mar 2013 0.048 0.075 nd apr 2013 0.032 0.057 nd may 2013 0.019 0.031 nd r2 oct 2011 0.000 0.000 0.764 may 2012 0.050 0.075 0.011 apr 2013 0.116 0.195 nd may 2013 0.050 0.070 1.526 r3 oct 2011 0.000 0.000 0.818 may 2012 0.007 0.012 0.063 na, not available; nd, not detected. no nco mm er cia l u se on ly m. stefanelli et al.82 duce some mc congener or other ctx. since w. naegeliana was simultaneously detected with other toxic species, no conclusive considerations on its toxicity have been reported so far. sivonen et al. (1990) reported that the neurotoxicity of bloom samples collected from finnish lakes was associated statistically with anabaena (dolichospermum) lemmermannii, anabaena (dolichospermum) flos-aquae and w. naegeliana. in other studies, various microcystins were found during blooms in which w. naegeliana was dominant (willame et al., 2005; baudin et al., 2006), and the presence of mcy genes in environmental samples characterized by the presence of this species was also reported (oberholster et al., 2006). however, up to now, there are no pure isolated cultures of w. naegeliana in the world cyanobacterial collection, and we also failed to maintain any culture after isolation. with respect to microcystis, and confirming our results on m. botrys, other studies in the field have observed significant correlations between the population abundances and mc concentrations (wang et al., 2010; horst et al., 2014, among others). the strong relationship between cb and mc in p. rubescens implies a high stability of population genetic structure over time (salmaso et al., 2014; kurmayer et al., 2016), which is indeed what we found in the flumendosa reservoir. but this is not always the case: in lake vico (central italy), where p. rubescens was dominant, the fraction of the mcyb+ cells was highly variable over time, especially during the blooms. in those conditions, the correlation between cell density and mc could explain only about 50% of mc variation (manganelli et al., 2016). the quantification of mcy genes of p. rubescens has been indicated as a better predictor of mc concentrations (ostermaier and kurmayer, 2010; hautala et al., 2013). however, in this study the application of the linear regression to the mcyb+ cells and mc reduced the percentage of mc variation explained, from >80% to ~50%, i.e., a value very close to what was observed by briand et al. (2008) in a population of p. agardhii and consistent with what we previously described in lake vico (manganelli et al., 2016). there are several possible explanations for this outcome, as summarized in kurmayer et al. (2016), such as other coexisting non identified toxic species, problems in the amplification of mcy genes or mutation inactivating mcy genes. or it could be also hypothesized that other factors, in addition to the genetic make up, influence the relationship between density and toxicity, and the relation between the whole population and mc, representing the environmental cell quota (ecq, sensu salmaso et al, 2014), includes those factors. since the relationship between mc content and the dynamics of the relative producer was quite strong, the slope of the linear regression could be a good estimate of the cell quota (salmaso et al., 2014), reducing the variability due to extreme values. a reliable estimate of the content of mc per cell can be very important when the possible risk for the population has to be assessed when only the density of cb is known. the ‘average’ density-based p. rubescens cell quota determined in this study were in the range of other values reported in literature (naselli-flores et al., 2007; briand et al., 2008; kosol et al., 2009; manganelli et al., 2010, 2016). interestingly, the cell quota determined for m. botrys were constantly about 10 times higher than p. rubescens and in the range of microcystis cell quota in the field estimated by yu et al. (2014) (0.001 and 1.326 pg cell–1 on the whole population and 0.012 and 1.876 pg cell–1 on the toxic cells). the high cell quota and the production of the more acutely toxic variants (mc-la and mc-lr, showing the lowest ld50 values) make m. botrys a highly toxic species, whose dynamic should be carefully monitored from spring to fall, for possible surface blooms. the levels of mc content during the study never exceeded 1 µg l–1 in all raw water discrete samples, excluding risk of acute and subchronic intoxication by drinking water for the population (who, 2004) and the domestic and wild animals (considering the action levels established by the california state epa, for dairy cows, beef, cattles and dogs; butler and linville, 2012). also the average cell quota was in the range of what is known for the cb species present. however, in the field some unpredictable events are possible (see may and december 2012) for which an exaggerated production of toxin per cell occurs (cell quota values up to 40 fold the average one in december) due to still unknown phenomena. the data used to compute these high values coincided with periods of high mc concentrations and p. rubescens density, making highly improbable the influence of measurement errors on the toxin quota estimates. a sudden increase in toxin concentration, in a time-frame of a few hours, has been observed in microcystis in the field and in laboratory experiment, and it has been associated to the fast increase in cell density and to faster mc production rate (wood et al., 2011, 2012). we speculate that the integrated sample could partly reproduce such a situation, since the ambient conditions of the cells change completely within few minutes, including a rapid increase of cell density. however, this explanation does not fit with the discrete december samples, as well as with the integrated r3 sample in may. this is definitely a field to explore, to define adequate strategies of management, since these situations are of particular concern when the risk for the exposed consumers has to be estimated. the relatively low cell densities found in the monitoring period is consistent with the low mc contamination level found in trout muscle. the discrepancy between the elisa and lc-ms/ms methods (estimated concentrations up to 0.9 ng g–1 and trace values, respectively) may no nco mm er cia l u se on ly cyanotoxins in a sardinian reservoir 83 be explained by multiple reasons: i) the matrix effects affecting the elisa detection, interfering with the optical density detection; ii) the elisa kit can reveal the presence of those mcs congeners for which analytical standard are not available for lc-ms/ms; and iii) the presence of conjugation products in fish tissue, which can cross-react with anti-mc antibodies. the trace levels detected by lc-ms/ms might be due to better compensation of matrix effects by the use of an appropriate calibration curve. however, low recovery of some mc variants could result to underestimation of mc concentrations in the muscle samples. in this respect, as pointed out by a review on the occurrence of cyanotoxins in food items (testai et al., 2016a), the availability of reliable, validated methods for detecting mc in complex matrices is a priority to collect good quality data to be used in the estimation of the possible risk due to human consumption. in addition, in view of the variability in the rate of toxins production by the cb communities, and in order to have a reliable estimate of the average ingestion of potentially contaminated fish and shell-fish grown in a specific water body, it is necessary to measure cyanotoxin content in the edible parts of representative species. furthermore, the sampling should be carried out over an adequate period of time to take into account seasonal variations in cyanotoxins production and possible differences among fish species depending on their diet, as well as on their capacity of retention/detoxication and elimination of cyanotoxin residues. it has indeed been shown that the bioaccumulation of mcs was species specific, with a tenfold higher bioaccumulation in catfish rather than in carp (singh and asthana, 2014). conclusions toxic cyanobacteria blooms can be a serious problem for the surface water distribution system in sardinia. in the oligo-mesotrophic flumendosa reservoir, toxic cyanobacteria were continuously present, although with relatively low densities and variations over time. three different species were alternatively dominating, namely p. rubescens (producing the demethylated form of mcrr and mc-lr), m. botrys (producing mc-lr and -la) and w. naegeliana (whose toxicity is still unknown). particular attention should be paid to the not visible deep p. rubescens summer blooms, since they are located were water is taken for drinking water supply. m. botrys, although not detectable during winter, was abundant in summer, meaning that inocula of the population are always present and should be carefully monitored since it produces the most toxic variants with a high cell quota. our results show that due to the relatively low density of cb and limited mc concentration found in the lake during the study period, there is no immediate risk of exposure to mc for the human and animal population. however, a variable cb community, characterised by the persistence of a significant toxic fraction, and the evidence of spotted variable toxin production rates suggest the need to take the system under control to manage possible sudden changes in toxins concentrations. the validity of the linear model applied to separate mc variants, to be further verified, show the possibility to expand its use in more complex cb communities, to obtain slopes representative of the studied systems. acknowledgments we thank dr roberta boi for helping in collecting field samples and dr valentina rosu for the work done. the work was supported by a grant from the ministero del lavoro, della salute e delle politiche sociali; dipartimento per la sanità pubblica veterinaria, la nutrizione e la sicurezza degli alimenti, contract number izs sa 04/10 rc. the funder had no role in study design, data collection and analysis, decision to publish, or preparation of the manuscript. the authors wish to thank the european cooperation in science and technology cost action es1105 cyanocost for networking and knowledge-transfer support. thanks to three anonymous reviewers who greatly improved the manuscript. references annadottér h, cronberg g, nystrand r, rylander r, 2005. endotoxins from cyanobacteria and gram-negative bacteria as the cause of an acute influenza-like reaction after inhalation of aerosols. ecohealth. 2:209-221. aktan y, lugliè a, sechi n, 2009. morphological plasticity of dominant species in response to nutrients dynamics in bidighinzu reservoir of sardinia, italy. turk. j. fish aq. sci. 9:137-144. baudin i, cagnard o, grandguillaume jj, do-quang z, 2006. algae and associated toxins & metabolites: methodology for risk assessment and risk management. water. pract. technol. 4:1-12. begliutti b, buscarinu p, marras g, sechi gm, sulis a, 2007. reservoirs water-quality characterization for optimization modelling under drought conditions. part i – reservoirs trophic state characterization, p. 239-261. in: g. rossi, t. vega and b. bonaccorso (eds.), methods and tools for drought analysis and management. springer, dordrecht. berry j, 2013. cyanobacteria toxins in food-webs: implications for human and environmental health, p. 531-589. in: a. rodriguez-morales (ed.), current topics in public health. botti p, dessena ma, fiori m, grillo sm, marcello a, matzuzzi c, pretti s, vacca s, 2001. the medio flumendosa reservoir. rendiconti seminario facoltà scienze università cagliari supplemento 71:209-220. briand e, gugger m, francois jc, bernard c, humbert jf, quiblier c, 2008. temporal variations in the dynamics of no nco mm er cia l u se on ly m. stefanelli et al.84 potentially microcystin-producing strains in a bloom-forming planktothrix agardhii (cyanobacterium) population. app. environ. microbiol. 74:3839-3848. briand jf, jacquet s, bernard c, humbert jf, 2003. health hazards for terrestrial vertebrates from toxic cyanobacteria in surface water ecosystems. vet. res. 34:361-377. briand jf, jacquet s, flinois c, avois-jacquet c, maisonnette c, leberre b, humbert jf. 2005. variations in the microcystin production of planktothrix rubescens (cyanobacteria) assessed from a four-year survey of lac du bourget (france) and from laboratory experiments. microb. ecol. 50:418-428. bucka h, wilk-woźniak e, 2002. a monograph of cosmopolitan and ubiquitous species among proand eukaryotic algae from water bodies in southern poland. karol starmach institute of freshwater biology, polish academy of sciences, kraków. buratti fm, manganelli m, vichi s, stefanelli m, scardala s, testai e, funari e, 2017. cyanotoxins: producing organisms, occurrence, toxicity, mechanism of action and human health toxicological risk evaluation. arch. toxicol. 91:1049-1130. buratti fm, scardala s, funari e, testai e, 2011. human glutathione transferases catalyzing the conjugation of the hepatoxin microcystin-lr. chem. res. toxicol. 24:926-933. butler ncj, linville r, 2012. toxicological summary and suggested action levels to reduce potential adverse health effects of six cyanotoxins. office of environmental health hazard assessment, california environmental protection agency, sacramento. cabras pa, rolesu s, orrù a, salati f, pintore a, garau p, funari e, lanni l, vodret b, 2013. [sviluppo di un modello di analisi del rischio di contaminazione da cianotossine in occasione di fioriture di cianobatteri tossici, ed eventuali riflessi sulle popolazioni animali domestiche e selvatiche e sull’ittiofauna di un lago della sardegna (lago alto flumendosa)].[report in italian]. ministero della salute; dipartimento per la sanità pubblica veterinaria, la nutrizione e la sicurezza degli alimenti. carbis cr, waldron dl, mitchell gf, anderson jw, mccauley i, 1995. recovery of hepatic function and latent mortalities in sheep exposed to the blue-green alga microcystis aeruginosa. vet. rec. 137(1):12-15. catherine a, quiblier c, yepremian c, got p, groleau a, vincon-leite b, bernard c, troussellier m, 2008. collapse of a planktothrix agardhii perennial bloom and microcystin dynamics in response to reduced phosphate concentrations in a temperate lake. fems microbiol. ecol. 65 (1):61–73. codd ga, edwards c, beattie ka, barr wm, gunn gj, 1992. fatal attraction to cyanobacteria? nature 359:110-111. dolman am, rücker j, pick fr, fastner j, rohrlack t, mischke u, wiedner c, 2012. cyanobacteria and cyanotoxins: the influence of nitrogen versus phosphorus. plos one 7:e38757. ernst b, neser s, o’brien e, hoeger sj, dietrich dr, 2006. determination of the filamentous cyanobacteria planktothrix rubescens in environmental water samples using an image processing system. harmful algae 5:281-289. frazier k, colvin b, styer e, hullinger g, garcia r, 1998. microcystin toxicosis in cattle due to overgrowth of blue-green algae. vet. hum. toxicol. 40:23-24. funari e, testai e, 2008. human health risk assessment related to cyanotoxins exposure. crit. rev. toxicol. 38:97-125. hallegraeff gm, anderson dm, cembella ad, 2004. manual on harmful marine microalgae monographs on oceanographic methodology 11. unesco, paris: 793 pp. hautala h, lamminmaki, u, spoof l, nybom s, meriluoto j, vehniainen m, 2013. quantitative pcr detection and improved sample preparation of microcystin-producing anabaena, microcystis and planktothrix. ecotoxicol. environ. saf. 87:49-56. hilborn ed, beasley val r, 2015. one health and cyanobacteria in freshwater systems: animal illnesses and deaths are sentinel events for human health risks. toxins. 71374-1395. holm-hansen o, lorenzen cj, holmes rw, strickland jdh, 1965. fluorometric determination of chlorophyll. j. cons. perm. int. explor. mer. 30:3-15. horst gp, sarnelle o, white jd, hamilton sk, kaul rb, bressie jd, 2014. nitrogen availability increases the toxin quota of a harmful cyanobacterium, microcystis aeruginosa. wat. res. 54:188-198. ibelings bw, chorus i, 2007. accumulation of cyanobacterial toxins in freshwater “seafood” and its consequences for public health, a review. environ. pollut. 150:177-192. isri, 2006. [le risorse idriche in sardegna. aggiornamento della valutazione intermedia].[in italian]. programma operativo della regione autonoma della sardegna.1. jespersen am, christoffersen k, 1987. measurements of chl-a from phytoplankton using ethanol as extraction solvent. arch. hydrobiol. 109: 445-454. komárek j, anagnostidis k, 1999. [cyanoprokaryota, 1: chroococcales]. in: h. ettl, g. gardner, h. heynig, d. mollenheuer (eds.), [süsswasserflora von mitteleuropa].[book in german]. gustav fischer, jena: 548 pp. komárek j, anagnostidis k, 2005. [cyanoprokaryota, 2: oscillatoriales]. in: b. büdel, l. krienitz, g. gärtner, m. schager (eds.), [süsswasserflora von mitteleuropa].[book in german]. elsevier, jena: 759 pp. komárek j, zapomělová e, 2007. planktic morphospecies of the cyanobacterial genus anabaena= subg. dolichospermum – 1. part: coiled types. fottea, olomouc 7:1-31. kosol s, schmidt j, kurmayer r, 2009. variation in peptide net production and growth among strains of the toxic cyanobacterium planktothrix spp. eur. j. phycol. 44:49-62. kurmayer r, deng l, entfellner e, 2016. role of toxic and bioactive secondary metabolites in colonization and bloom formation by filamentous cyanobacteria planktothrix. harmful algae 54:69-86. loizzo a, sechi n, volterra l, contu a, 1988. some features of a bloom of oscillatoria rubescens d.c. registered in two italian reservoirs. wat. air soil pollut. 38:263-271. lopez rodas v, costas e, 1999. preference of mice to consume microcystis aeruginosa (toxin-producing cyanobacteria): a possible explanation for numerous fatalities of livestock and wildlife. res. vet. sci. 67:107-110. lugliè a, gesuina l, manca b, sechi n, 1997. [studi limnologici sul lago alto flumendosa (sardegna centrale): stato trofico e fitoplancton].[article in italian]. boll. soc. sarda sci. nat. 31:83-100. lugliè a, padedda bm, bazzoni am, caddeo t, casiddu p, farina p, lai g, manca b, mariani ma, pulina s, satta ct, stacca d, sechi n, 2013. [aspetti dell’ecologia dei sistemi acquatici della sardegna e loro principali problematiche]. [article in italian]. proceedings assemblea generale e del no nco mm er cia l u se on ly cyanotoxins in a sardinian reservoir 85 symposium i mari delle isole (resèaux d’excellence des territoires insulaires – reti): 8 pp. manganelli m, scardala s, stefanelli m, vichi s, mattei d, bogialli s, ceccarelli p, corradetti e, petrucci i, gemma s, testai e, funari e, 2010. health risk evaluation associated to planktothrix rubescens: an integrated approach to design tailored monitoring programs for human exposure to cyanotoxins. wat. res. 44:1297-1306. manganelli m, stefanelli m, vichi s, andreani p, nascetti g, scialanca f, scardala s, testai e, funari e, 2016. cyanobacteria biennal dynamic in a volcanic mesotrophic lake in central italy: strategies to prevent dangerous human exposures to cyanotoxins. toxicon 115:28-40. marchetto a, padedda bm, mariani ma, lugliè a, sechi n, 2009. a numerical index for evaluating phytoplankton response to changes in nutrient levels in deep mediterranean reservoirs. j. limnol. 68:106 121. mariani ma, padedda bm, kaštovský j, buscarinu p, sechi n, virdis t, lugliè a, 2015. effects of trophic status on microcystin production and the dominance of cyanobacteria in the phytoplankton assemblage of mediterranean reservoirs. sci. rep. 5:1-16. meregalli m, trebeni f, manca b, lugliè a, 2002. [stato trofico e fitoplancton nel lago alto flumendosa].[article in italian]. atti associazione italiana oceanografia limnologica 15:87-96. messineo v, mattei d, melchiorre s, salvatore g, bogialli s, salzano r, mazza r, capelli g, bruno m, 2006. microcystin diversity in a planktothrix rubescens population from lake albano (central italy). toxicon 48:160-174. messineo v, bogialli s, melchiorre s, sechi n, lugliè, antonella gl, casiddu p, mariani ma, padedda bm, di corcia a, mazza r, carloni e, bruno m, 2009. cyanobacterial toxins in italian freshwaters. limnol. ecol. manag. inl. waters 39:95-106. moreira c, vasconcelos v, antunes a, 2013. phylogeny and biogeography of cyanobacteria and their produced toxins. mar. drugs 11:4350-4369. naselli-flores l, barone r, chorus i, kurmayer r, 2007. toxic cyanobacterial blooms in reservoirs under a semiarid mediterranean climate: the magnification of a problem. environ. toxicol. 22:399-404. oberholster pj, botha a-m, cloete te, 2006. toxic cyanobacterial blooms in a shallow, artificially mixed urban lake in colorado, usa. lakes & reservoirs 11:111-123. ostermaier v, kurmayer r, 2009. distribution and abundance of nontoxic mutants of cyanobacteria in lakes of the alps. microb. ecol. 58:323-333. ostermaier v, kurmayer r, 2010. application of real-time pcr to estimate toxin production by the cyanobacterium planktothrix sp. appl. environ. microbiol. 76:3495-3502. paerl hw, paul vj, 2012. climate change: links to global expansion of harmful cyanobacteria. wat. res. 46:1349-1363. papadimitriou t, kagalou i, leonardos id, 2012. seasonally accumulation of microcystins in the various tissues of an endemic and protected fish species (rutilus panosi) with different sizes. clean soil air water 40:402-407. pellegrini s, grilli caiola m, sechi n, 1995. microcystis from lake liscia (sardinia italy). paper presented at the 8th viii int. symp. phototrophic prokaryotes abstract, urbino. planas d, paquet s, 2016. importance of climate change-physical forcing on the increase of cyanobacterial blooms in a small, stratified lake. j. limnol. 75:201-214. pulina s, padedda bm, sechi n, lugliè a, 2011. the dominance of cyanobacteria in mediterranean hypereutrophic lagoons: a case study of cabras lagoon (sardinia, italy). scientia marina 75:111-120. rajaniemi-wacklin p, rantala a, mugnai ma, turicchia s, ventura s, komárková j, lepistö l, sivonen k, 2005. correspondence between phylogeny and morphology of snowella spp. and woronichinia naegeliana, cyanobacteria commonly occurring in lakes. j. phycol. 42:226-232. rantala a, rajaniemi-wacklin p, lyra c, lepistö l, rintala j, mankiewicz-boczek j, sivonen k, 2006. detection of microcystin-producing cyanobacteria in finnish lakes with genus-specific microcystin synthetase gene e (mcye) pcr and associations with environmental factors. appl. environ. microbiol. 72:6101-6110. reynolds cs, jaworsky ghm, cmiech ha, leedale gf, 1981. on the annual cycle of the blue-green algae microcystis aeruginosa kütz. emend. elenkin. philos. t. r. soc. b 293:419-477. rippka r, 1988. isolation and purification of cyanobacteria. method enzymol. 167:3-27. saker ml, jungblut ad, neilan ba, rawn dfk, vasconcelos vm, 2005. detection of microcystin synthetase genes in health food supplements containing the freshwater cyanobacterium aphanizomenon flos-aquae. toxicon 46:555-562. salmaso n, copetti d, cerasino d, shams s, capelli c, boscaini a, valsecchi a, pozzoni f, guzzella l, 2014. variability of microcystin cell quota in metapopulations of planktothrix rubescens: causes and implications for water management. toxicon 90:82-96. sechi gm, sulis a, 2009. water system management through a mixed optimization-simulation approach. j. wat. resour. plan. manag. 135:160-170. sechi n, lugliè a, 1992. limnological studies on man-made lakes in sardinia (italy). mem. ist. it. idrobiol. 50:365-381. sechi n, lugliè a, 1996. phytoplankton in sardinian reservoirs. g. bot. ital. 130:977-994. sedmak b, eleršek t, grach-pogrebinsky o, carmeli s, sever n, lah tt, 2008. ecotoxicologically relevant cyclic peptides from cyanobacterial bloom (planktothrix rubescens) – a threat to human and environmental health. radiol. oncol. 42:102-113. singh s, asthana, r, 2014. assessment of microcystin concentration in carp and catfish: a case study from lakshmikund pond, varanasi, india. bull. environ. contam. toxicol. 92:687-692. sivonen k, niemelä si, niemi rm, lepistö l, luoma th, räsänen la, 1990. toxic cyanobacteria (blue-green algae) in finnish fresh and coastal waters. hydrobiologia 3:256-275. stewart i, seawright aa, shaw gr, 2008. cyanobacterial poisoning in livestock, wild mammals and birds – an overview, p. 611-637. in: h.k. hudnell (ed.), cyanobacterial harmful algal blooms state of the science and research needs. springer, dordrecht. sulis a, buscarinu p, soru o, sechi gm, 2014. trophic state and toxic cyanobacteria density in optimization modeling of multi-reservoir water resource systems. toxins (basel) 6:1366-1384. no nco mm er cia l u se on ly m. stefanelli et al.86 testai e, buratti fm, funari e, manganelli m, vichi s, arnich n, biré r, fessard v, sialehaamoa a, 2016a. review and analysis of occurrence, exposure and toxicity of cyanobacteria toxins in food. efsa supporting publication 2016: en-998. 309 pp. available at: http://www.efsa.europa.eu/sites/default/files/ scientific_output/files/main_documents/998e.pdf testai e, scardala s, vichi s, buratti fm, funari e, 2016b. risk to human health associated with the environmental occurrence of cyanobacterial neurotoxic alkaloids anatoxins and saxitoxins. crit. rev. toxicol. 29:1-35. utermhöl h, 1958. [zur vervollkhung der quantitativen phytoplanktonmethodik].[article in german]. mitt. int. verein. theor. angew. limnol. 9:1-38. vaitomaa j, rantala a, halinen k, rouhiainen l, tallberg p, mokelke l, sivonen k, 2003. quantitative real-time pcr for determination of microcystin synthetase gene e copy numbers for microcystis and anabaena in lakes. appl. environ. microbiol. 69:7289-7297. vichi s, lavorini p, funari e, scardala s, testai e, 2012. contamination by microcystis and microcystins of blue-green algae food supplements (bgas) on the italian market and possible risk for the exposed population. food chem. toxicol. 50:4493-4499. walsby ae, avery a, 1996. measurement of filamentous cyanobacteria by image analysis. j. microbiol. meth. 26: 11-20. wang q, niu y, xie p, chen j, ma z, tao m, qi m, wu l, guo l, 2010. factors affecting temporal and spatial variations of microcystins in gonghu bay of lake taihu, with potential risk of microcystin contamination to human health. thescientificworldjo 10:1795-1809. whitton ba (ed.), 2012. ecology of cyanobacteria ii. their diversity in space and time. springer, dordrecht: 760 pp. who, 2004. guidelines for drinking water quality. 1. recommendations. world health organization, geneva. willame r, jurczak t, iffly jf, kull t, meriluoto j, hoffmann l, 2005. distribution of hepatotoxic cyanobacterial blooms in belgium and luxembourg. hydrobiologia 551:99-117. wood r, 2016. acute animal and human poisonings from cyanotoxin exposure a review of the literature. environ. int. 91:276-282. wood sa, dietrich dr, cary sc, hamilton dp, 2012. increasing microcystis cell density enhances microcystin synthesis: a mesocosm study. inland waters 2:17-22. wood sa, rueckert a, hamilton dp, cary sc, dietrich dr, 2011. switching toxin production on and off: intermittent microcystin synthesis in a microcystis bloom. environ. microbiol. rep. 3:118-124. yu g, zhu m, li r, tan w, jiang y, song g, 2014. variation of microcystis and microcystins coupling nitrogen and phosphorus nutrients in lake erhai, a drinking-water source in southwest plateau, china. environ. sci. pollut. res. int. 21:9887-9898. no nco mm er cia l u se on ly layout 1 introduction estuaries are among the most biologically productive and economically valuable ecosystems in the world. the hydrological regime and geomorphological features of estuaries make such ecosystems good shelters, feeding and breeding grounds for many marine fish species at various stages of their life cycles (beck et al., 2001; pihl et al., 2002; harrison and whitfield, 2006; kostecki et al., 2010; selleslagh et al., 2015). estuaries provide essential and complex habitats to sustain different freshwater and marine fish stocks. the estuarine complexity provides juvenile fishes with adequate protection from predation while attaining high growth rates supported by a high primary productivity (kostecki et al., 2010; whitfield, 2016). nevertheless, the ecological status of most estuaries is threatened due to modifications made by anthropogenic activities. estuaries are subjected to increasing human pressure through urbanization, overfishing and upstream catchment disturbances resulting from deforestation, water abstraction for reservoir or multi-dams construction and water diversion for irrigation activities (abrantes et al., 2014). the physical, chemical and biological characteristics of estuaries are dynamic both on temporal and spatial scales (selleslagh et al., 2015). the main drivers of such fluctuations are hydrological regimes, encompassing freshwater discharges from the upstream, coastal wave activity and tidal water flow at the mouths of the estuaries (gillson, 2011; hoeinghaus et al., 2011). the interference of such hydrological regime such as upstream water abstraction for irrigation and damming for hydropower generation causes retention and thus reduction of nutrients, sediment (that serve as an ecological niche to some species), quantity and quality of organic matter reaching the estuary (han et al., 2016). variations in the hydrological regime impact habitat availability and diversity, primary production, trophic interactions and consequently prompt responses in food web structure and functioning (gillson, 2011; whitfield, 2016; donázar-aramendía et al., 2019). more so, the regulation of the flow rate of freshwater reaching the estuaries greatly influences the diversity of basal organic matter fueling the estuarine food webs (abrantes et al., 2013). juvenile marine fish with predatory behavior constitute a great part of estuarine fish communities. they are highly targeted by local or artisanal fisheries while still at juvenile and sub-adult stages, as they attain large body sizes before reaching maturity (shaw et al., 2016; mwijage et al., 2018a). as such, they have lucrative market price when compared to the small fish that feed at lower trophic levels. predatory fish and other higher order consumers coexist article diet and isotopic metrics of predatory and prey fish in two estuaries with different degrees of anthropogenic disturbances: the case study of wami and pangani rivers in tanzania theresia john lyasenga1, alistidia p. mwijage2, dativa joseph shilla3, john andrew marco mahugija4, lydia gaspare1, daniel abel shilla1, prosper laurent mfilinge1 1school of aquatic sciences and fisheries technology, university of dar es salaam; 2tanzania fisheries research institute (tafiri), dar es salaam; 3department of chemistry, dar es salaam university of college of education (duce); 4college of natural and applied sciences, department of chemistry, university of dar es salaam, tanzania abstract diverse anthropogenic activities including alteration of hydrological regime and agricultural development in the upstream of the river catchments modify the structural components of communities and ecological roles of species in estuarine ecosystems. the present study compared the diet, carbon (δ13c) and nitrogen (δ15n) isotope ratios and layman community metrics of predatory fish and their prey-fish between two estuaries with different degrees of anthropogenic disturbances. the layman community metrics were estimated following the bayesian approach. it was revealed that the diets of the predators arius africanus, epinephelus coioides, sillago sihama and pomadasys argenteus, and their isotopic values including their main prey, the fish valamugil buchanani, were significantly different between wami and pangani estuaries (permanova, pseudo-f≥3.1, p=0.05). the comparison test of isotopic niche width at estuary level showed that the isotopic niche of wami estuary was slightly narrower (3.906.94 at 95% ci) than that of pangani (5.70-9.69 of 95% ci). the comparison of the δ15n values and range of species between estuaries suggest shifts in trophic position of the species in wami estuary relative to that of pangani. furthermore, the layman metric indices of trophic diversity and redundancy of wami estuary were substantially smaller, when compared to that of pangani estuary. these findings indicate that, even though the wami estuary stands under a conservation status, the extent of disturbances in the upstream is potentially high and enough to induce the ecological changes at the base of the food web downstream, giving rise to subsequent effects transmitted to higher-order consumers. as a result, the ecological redundancy and ecosystem complexity of wami are possibly compromised relative to that of pangani estuary. no nco mm er cia l u se on ly t.j. lyasenga et al.84 in estuaries as they exert reciprocally positive indirect effects by regulating the populations of their respective preys, and in that way, prevent competitive exclusion of each other’s preys (sanders et al., 2018). they also link pelagic, benthic and detritivore food webs in estuarine ecosystems through predator-prey interactions (rooney and mccann, 2012). because of the array of ecological roles they play, by studying these predators, one can get insights into the role of predatory fish for maintaining the structure and functioning of estuarine ecosystem. the pangani and wami river estuaries are situated about 55 kilometers apart in the coastal northern part of tanzania. the pangani estuary is managed by pangani water basin office authority, and thus it is an open access estuary. the wami estuary is part of saadani national park (sanapa) and is therefore a protected estuary but the whole river is managed by wami-ruvu water basin office authority. however, the upstream part of pangani and wami rivers are subjected to different anthropogenic activities that may be affecting the ecology of estuarine fish and other biota in different ways. in the upstream portion of pangani catchment, water is mainly abstracted by multi-damming for hydroelectric power generation and irrigation (shaghude, 2016; mwijage, et al., 2018a). the retained sediments and organic particles in the reservoirs and the elevated load of nutrient inputs arising from these activities affect different ecological processes downstream the rivers including estuaries. yet, water upstream of wami estuary is widely used for fishing and irrigation in rice paddies and commercial sugarcane plantations (eeden and koppen, 2016). there is an indication that the multi-damming activities in pangani catchment have reduced the nutrients loading downstream (selemani et al., 2017) and caused some predatory fishes to largely rely on the estuarine and marine organic matter sources (mwijage et al., 2018a). in the wami estuary, the basal primary food sources sustaining the components of the estuarine food webs is unknown. despite the fact that the wami estuary receives elevated nutrients and suspended particles from the upstream, according to kiwango et al. (2015), the estuary is still ecologically healthy. generally, the anthropogenic disturbances of tanzanian estuaries and coastal ecosystem is high (see, e.g., semba et al., (2016) and shaghude (2016)), but little is known about the implications of such disturbances on the trophic niche, and other related dynamics, of fish assemblage and, consequently, on the entire estuarine ecosystem. this paper aims at getting more insight into the abovementioned perspectives, by comparing the diet and other trophic dynamics of fish species at the different degrees of anthropogenic disturbances present in wami and pangani estuaries. stable isotopes and stomach content analyses are employed in complementarity to describe the trophic structure in numerous ecosystems. stable isotopes ratios of carbon (δ13c) and nitrogen δ15n can be used as a robust tool to evaluate changes in the diet of native species in relation to the expansion of invasive species (garcía et al., 2020). stable isotopes are also used to analyze the trophic niche, dietary overlap and other trophic dynamics of food web components across varied ecosystems (layman et al., 2012; jackson and parnell, 2016; gutmann and britton, 2018; wang et al., 2018; zhang et al., 2019). through the bayesian framework, layman metrics are estimated to enable evaluation of the numerous human-induced impacts on food webs, including ecological redundancy, resilience, diversity and trophic stability of ecosystems (rooney and mccann, 2012; matich et al., 2017; donázar-aramendía et al., 2019). in this study, we examine the degree of dietary and isotopic niche flexibility of predatory fish and their phytodetritivore prey-fish in the adjacent estuaries of wami and pangani rivers. specifically, our study intended to (i) assess the extent of similarity or overlaps in diet of predatory fish; ii) estimate isotopic values and trophic niche of predatory fish and their phyto-detritivore prey-fish populations; and (iii) elucidate the status of complexity, resilience and redundancy of two estuarine ecosystems using layman metrics. in presenting and discussing our results, we provide critical information to further understand the effects of the anthropogenic activities on fish communities and the resulting impacts on the tanzanian river ecosystems as well as refine management approaches for sustainability of riverine and coastal resources in general. methods study area the study sites were based in the two macro-tidal estuaries of semi-diurnal type adjacent to each other, located in the northern coastal part of tanzania, namely, wami and pangani estuaries (figure 1). both estuaries are permanently open, connected to the sea. the wami estuary is part of the saadani national park (sanapa) which was established in 2005 and thus, it is strictly protected. the tidal range in this estuary reaches up to 4 m during spring tides and the influence may extend up to 8 km upstream (kiwango et al., 2018a). during the wet season, the freshwater plume of wami river extends for several hundred meters into the indian ocean during both low and high tides. only the first five kilometers from the shore are occupied by mangroves. the estuary shows turbidity ≥400 ntu and suspended sediment concentrations of <100 mg l-1 during the dry season, and >800 mg l-1 during the wet season (kiwango et al., 2018a). a salt wedge is common in the offshore areas of wami, several hundred no nco mm er cia l u se on ly estuarine fish trophic dynamics and anthropogenic disturbances 85 meters from the mouth of the river in a northeasterly direction. the pangani estuary is located further north of wami estuary (figure 1) with a wave amplitude of about 3.5 m during the spring tides and 3.0 m at neap tides (mwijage et al., 2018a). the estuary extends to about 20 km from the estuary mouth (mwijage et al., 2018a). it has an extensive mangrove forest interspaced with coconut/palm trees. pangani estuary experiences strong mixing of fresh and saline water. the hydrological regime is mainly regulated by the multi-reservoirs upstream of the catchment. moreover, both estuaries serve as breeding and nursery grounds for marine fish, prawns, and birds, and they are known as the most productive areas of prawns for tanzania. wami estuary, to a larger extent, hosts a larger number of hippopotamus and crocodiles compared to those occurring in the pangani estuary. field sampling procedures of fish fish samples were collected in wami and pangani river estuaries in july 2019 during the daytime. july is a dry-season month and within the peak of the southeast monsoon. thus, field sampling was conducted after the peak of the rainy season which usually occurs in may; this allowed to reduce seasonal biasness by capturing isotopic values that would reflect dietary sources that were consumed by fish during the wet and dry seasons. this is based on the premise that isotopic values of consumer tissues indicate their long-term assimilated diet of at least three months. fish species were caught by using monofilament gill nets of multiple mesh sizes and seine net with the dimensions of 15 m length, 1.5 m width and mesh size 0.5 mm. from the total landing, five abundant predatory fish species, namely arius africanus (günther 1867), epinephelus coioides (hamilton, 1822), sillago sihama (forsskål, 1775) and pomadasys argenteus (forsskål, 1775), and their main potential prey such as the phyto-detritivore valamugil buchanani (bleeker, 1853), were selected to study their trophic niches flexibility. the selection was based on what predatory fish species of marine affinity are abundant throughout the seasons (bianchi, 1995; fiu-glows, 2016). notably, marine fish species inhabiting wami estuary include a. africanus, v. buchanani, chanos chanos, ambassis gymnocephalus, hilsa kelee, thryssa spp., leiognathus equulu, pomadasys spp., epinephelus spp. and tetraodontidae, (anderson et al., 2007; fiu-glows, 2016). however, the diversity and abundance of these species decrease sharply as the distance from the mouth of the river to the upstream increases (fiu-glows, 2016). in pangani estuary, fish species with a marine origin commonly seen include c. chanos, a. gymnocephalus, gerres filamentosus, v. buchanani, lutjanus argentimaculatus, sillago sihama, epinephelus spp. and species from the families of sphyraenidae, hemiramphidae, gobiidae and carangidae (pwbo/iucn, 2008; bianchi, 1995; mwijage et al., 2018a). during the sampling, the predatory fish species e. coioides and a. africanus were caught in both estuaries along with their main prey, i.e., v. buchanani. both s. sihama and p. argenteus were abundantly caught in pangani estuary, while p. argenteus was greatly caught within the plume setting at short distance from the mouth of the wami estuary. since the two above-mentioned species are classified as zoobenthivore under the feeding mode functional grouping (elliott et al., 2007), their trophic requirements are equivalent. due to that, we decided to select them as another pair of comparison between estuaries. to get enough samples for stomach content analysis, additional samples of p. argenteus were bought from the local market immediately after being landed by local fishermen. all fish species caught were stored in ice boxes containing dry ice immediately after collection. later, total length and weight were recorded, and all fish samples frozen at -20°c on the same day and transported to the figure 1. map of wami and pangani river estuaries showing the study area. no nco mm er cia l u se on ly t.j. lyasenga et al.86 laboratory at the university of dar es salaam, college of agricultural sciences and fisheries technology for further analyses. fish stomach content analysis in the laboratory, stomachs of all fish species caught were extracted and subjected to content analysis. the fish stomachs were dissected under a microscope to enable the determination of the frequency of occurrence (%fo) and percentage volumetric contribution of each dietary item to the total stomach volume (%v). this was done following the point method described by hyslop (1980). the diet items were identified at the possible low taxonomic level. each food content in the stomach was allotted one of the number of points 0, 1, 2, 4, 8 and 16 based on its volume, from the smallest to the largest value, respectively. then percentage volumes within each stomach analyzed were then calculated to give the percentage composition in a diet of individual fish species. also, vacuity index or the stomach emptiness index (vi) of each species was estimated using the equation given by euzen (1987): vi = (the number of empty stomachs of species i /total number of the stomachs of species i examined) × 100. by using the dietary volume data, the standardized levin’s (1968) diet niche breadth (ba) was estimated for each species in all estuaries following hurlbert (1978). the following formula was used: where ba is the standardized trophic niche breadth, pij is the proportion of food category j in the diet of the species i and n = total number of food items in the diet of species i. the ba values range between 0, when only one resource is used, and 1, when all resources are used equally. in addition to that, the dietary overlap (o) of the predatory fish species in each estuary were estimated by using pianka’s index of niche overlap (pianka, 1973), which considers that the different dietary resources are equally accessible to all species. the formula used was: where pij and pik are the proportions of diet category i comprised in the diets of j and k, respectively. both diet niche breadth (ba) and overlap (o) were considered low when the values fall within 0-0.39 range, intermediate when they are within 0.4-0.6, and high when ranged from 0.61-1. overall interspecific variations in the diet of the fish species between and within the two estuaries were assessed by main and pairwise permutation multivariate analysis of variance (permanova). the multivariate analysis, including cluster analysis that produces dendrograms to measure the similarity in diet, and the permanova tests were run based on the bray curtis similarity matrices made from the square root transformed volumetric percentage dietary data using primerv6 with permanova + add-on module statistical packages (anderson et al., 2008). before conducting any statistical test, diet data were subjected to normality and homogeneity of variance tests, the assumptions were not met. thus, nonparametric method was opted and because of the multiple variables, the multivariate analyses of variance tests were selected. stable isotopes and trophic niche indices analyses for each fish species collected for stomach content analysis, representative samples were selected randomly to extract dorsal white muscle tissues for stable isotope analysis (sia). the sample size (n) and length size (total length tl in cm) of each species were different: a. africanus from wami = tl ranged 16.6-29.9 cm (n=16); a. africanus of pangani = tl ranged 13.5-44.2 cm (n=16); e. coioides of wami = tl ranged 15.3-30.5 cm (n=6); e. coioides of pangani = tl ranged 15.6-32.0 cm (n=6); p. argenteus of wami = tl ranged 31.3-41.0 cm (n=5); and s. sihama of pangani = tl ranged 15.0-25.6 cm (n=8). the extracted muscle tissues were oven-dried at 60°c and ground to a powder form. a subsample of each sample was weighed to the nearest 0.9 mg and placed into tin capsules for δ13c and δ15n analysis. the δ13c and δ15n compositions of each sample were determined simultaneously using a dual pumped sercon model 20-20 continuous flow isotope ratio mass spectrometer (cf/irms or ea/irms) linked to a thermo model ea1110 elemental analyser (ea). the equipment utilizes dual reaction tubes (combustion/reduction), a magnesium perchlorate drier and a carbosieve g separation column. stable isotope results were expressed in the delta notation (δ) relative to the known standards, vienna pee dee belemnite (vpdb) for 13c/12c and atmospheric nitrogen for 15n/14n ratios. the δ13c or δ15n were calculated using the following formula: δ = [(rsample/rstandard) – 1] x 1000 where r is either 13c/12c or 15n/14n. these isotope analyses were carried out at the oea labs limited in the united kingdom. before statistical analysis of isotopic results, these latter were evaluated to gauge if the lipid contents were likely to affect the analysis using the c:n ratio (post et al., 2007). the c:n ratios of all species were below 3.5, thus, in accordance with post et al. (2007), the levels of lipids in tissues of fish samples were not enough to skew isotopic results. to test the variations in isotopic values among fish species and between estuaries, the isotope data were subjected to main and pairwise permanova tests. two fixed factors, no nco mm er cia l u se on ly estuarine fish trophic dynamics and anthropogenic disturbances 87 species and estuaries, were considered. the tests were run based on the euclidian distance resemblance matrix made from the normalized isotopic data. furthermore, to enable quantitative comparisons of the trophic niche of fish communities within and between the two estuaries, the standard ellipse area corrected for small sample size (seac), total convex hull area (ta), and other five layman metrics were estimated using the stable isotope bayesian ellipses in r (siber, jackson et al., 2011) package version 4.0.5 (r core team, 2021). standard ellipse area corrected for small sample size and ta and were estimated as quantitative proxies of the isotopic niche width that measure total trophic diversity or niche area. however, these two metrics differ, because seac is less sensitive to outliers and sample size, as it is corrected for small sample sizes, than the values of ta. for statistical comparison of seac between the communities investigated, the posterior distribution of the bayesian standard ellipse (seab) and the 95% credible intervals (cis) were calculated (jackson et al., 2011). to test whether the isotopic niche space of fish population of wami estuary is larger, smaller or similar to the comparable population of pangani, the 95% ci of seab and ta of those targeted pairs of fish species (see previous paragraphs) were compared. when the ci of distinct pairs of species did not completely overlap, the trophic niche of the comparable populations was significantly different. by following jackson et al. (2011), the probability that seab or ta of species of wami was smaller than that of pangani was also quantified. the proportions of overlaps for the comparable groups were also calculated using bayesian overlap and compared as they had to match with the density plot for the 95% prediction ellipsoids. the overlaps estimated aimed at describing how much of each species’ niche overlapped with others between and within the estuaries. as per layman et al. (2007), other metrics estimated using siber package were δ13c range (cr), δ15n range (nr), mean distance to the centroid (cd), mean nearest neighbor distance (m-nnd) and standard deviation (sd) of the m-nnd (sd-nnd). these bayesian-derived estimates allow statistical comparisons between communities without any restriction on the number of the groups within the communities (layman et al., 2012; donázar-aramendía et al., 2019). the cr was used to indicate trophic diversity or niche diversification at the base of the food web. the nr was used to infer both trophic diversity and the trophic length, meaning that a larger range infers a wide trophic spectrum. the cd, which is the average euclidian distance of each community component to the centroid, was used to infer the average degree of trophic diversity. the two metrics, m-nnd and sd-nnd were used to evaluate to what extent the trophic redundancy of the fish community vary between the estuaries. the mnnd measures the density of species packing, meaning that species or groups with similar trophic ecologies show smaller values of m-nnd and thus represent increased trophic redundancy. the sd-nnd metric indicates the evenness of species packing. this means that lower sdnnd is indicative of more even distribution of species, the indictor of higher trophic redundancy as a result of different groups having more similar trophic ecologies (layman et al., 2007). the sd-nnd also show diversification of trophic niches (donázar-aramendía et al., 2019). results diet compositions of predatory fish a total of 420 stomachs of four estuarine predatory fish species were examined. out of that, 332 (79%) contained food items that were used for further analyses. irrespective of where species were caught, seven dietary items were identified in the stomachs of predatory fish species, namely: prey fish, stomatopods, bivalves and other unidentified mollusks, prawns, shrimps, polychaetes and crabs. digested materials and remains of hard parts of preys were also seen in some species. detrital materials and sand particles were observed as accidental items ingested by a. africanus. at the species level, the values of vacuity index (vi) were considerably high in both estuaries (table 1) and the patterns of vi showed that the predatory fish species were more voracious in wami relative to that of the pangani estuary. with the exclusion of polychaetes observed in the stomachs of both a. africanus and s. sihama, as well as detrital materials found only in the stomachs of a. africanus, all other preys were common in all four predatory fish species under investigation, albeit with different volumetric contribution (table 1). arius africanus of wami estuary fed mainly (by volume) on bivalves (27%) followed by other mollusks (26%) and small fish (19%), while a. africanus of pangani estuary ingested large amounts of bivalves (36%), other mollusks (31%) and crabs (9%). stomatopods and shrimp prey were rarely seen in the stomachs of a. africanus from wami estuary. fish prey, prawns and crabs were the three most consumed preys equally by e. coioides in the wami estuary, whereas prawns, fish, and stomatopods occupied large percentage of diet for individuals caught in wami estuary (table 1). the diet of p. argenteus of wami estuary was mainly dominated by prawns and to some extent by fish and mollusks, while s. sihama of pangani estuary ingested large amounts of fish, shrimps, and substantial amounts of polychaetes (table 1). despite grouping the prey types at higher taxonomic level, it was noticed that valamugil spp. were dominant among the fish preys encountered in the stomachs of all the fish examined. no nco mm er cia l u se on ly t.j. lyasenga et al.88 multivariate analyses of dietary composition in predatory fish the dendrogram prepared based on the volume of prey items showed that the diet items consumed by the population of a. africanus of wami and that of the pangani estuaries were highly similar for about 75% (figure 2). it was also revealed that about 50% of the diet of e. coioides of wami was shared with p. argenteus from the same estuary (figure 2). the prey items consumed by e. coioides population of wami were to some extent different (about 40% similarity) to that of the population of pangani estuary. the results further highlighted that the preys consumed by pangani-drawn e. coioides and s. sihama were far related (40% figure 2. dendrogram for hierarchical clustering of dietary composition (by percentage volume) in arius africanus, epinephelus coioides, pomadasys argenteus and sillago sihama collected from pangani (p) and wami (w) estuaries. table 1. diet composition of arius africanus, epinephelus coioides, pomadasys argenteus and sillago sihama found in their stomachs and dietary indexes of each prey item: frequency of occurrence (%f), percentage volumetric contribution (%v), number of stomachs with food/content (n) and vacuity index percentage (%vi). food items wami estuary pangani estuary a. africanus e. coioides p. argenteus a. africanus e. coioides s. sihama (n = 48) (n = 52) (n = 54) (n = 74) (n = 52) (n = 52) %fo %v %fo %v %fo %v %fo %v %fo %v %fo %v fish 50.0 19.4 51.6 22.4 31.1 11.6 20.8 7.6 64.7 23.2 45.7 33.7 stomatopods 2.4 1.2 51.6 11.9 13.3 4.0 64.7 19.4 2.9 2.6 bivalves 42.9 27.0 9.7 1.5 26.7 9.0 58.3 35.7 other mollusks 40.5 26.0 26.7 10.1 62.5 31.2 35.3 8.8 22.9 13.1 prawns 51.6 22.1 84.4 54.1 16.7 4.8 70.6 32.1 polychaetes 19.1 5.7 20.8 6.5 22.9 17.1 crabs 35.7 12.0 64.5 22.4 26.7 8.0 31.3 9.3 35.3 12.1 5.7 1.5 shrimps 14.3 3.6 41.9 14.8 11.8 4.4 28.6 20.6 sand/detritus 4.8 3.0 2.1 1.0 hard tissues/parts 19.4 3.9 2.2 0.9 6.3 1.9 unidentifiable/digested 18.1 2.1 6.5 1.0 15.6 2.3 14.6 1.9 25.7 11.4 vacuity index (vi) 9.76 21.21 15.63 22.58 33.33 23.56 no nco mm er cia l u se on ly estuarine fish trophic dynamics and anthropogenic disturbances 89 average similarity) to each other. the low levels of similarity in prey consumed by fish were also seen among the e. coioides, a. africanus from the two estuaries and p. argenteus drawn from wami estuary (40% average similarity, figure 2). the results of permanova test paired with that of dendrogram showed significant differences among species (pseudo-f=21.0; p=0.001; table 2) as well as the interactions between main factors, species, and estuaries (pseudo-f=21.0; p=0.001; table 2). even the pairwise permanova tests indicated spatial intra-specific diet variations of each of the three pairs of predatory fish species caught from wami and pangani estuaries (t≥1.88; p≤0.05; table 3). dietary niche breadth and overlaps of predatory species the values of index of dietary niche breadth (ba) of the predatory fish species varied from 0.29 to 0.7 (figure 3). despite the dietary volume of e. coioides varied between estuaries (table 2), the species presented much higher values of ba as compared to other species (figure 3). arius africanus from wami showed moderate niche breadth, whereas specimens from pangani estuary showed lower values of ba (figure 3). a different situation was noticed for the comparable pair of p. argenteus from wami and s. sihama from pangani where the former species showed lower, and the latter species showed moderate dietary niche breadths (figure 3). furthermore, the analysis of dietary overlaps based on pianka’s index (o) revealed much high values (o>0.6) for the following pairs: a. africanus of wami and a. africanus of pangani, e. coioides of wami and e. coioides of pangani, as well as p. argenteus of wami and e. coioides of pangani, (table 4) which is an indication of low trophic flexibility of these species in two estuaries. however, the lowest index of overlap (o=0.2) was presented by the pair of p. argenteus of wami versus s. sihama of pangani (table 4). variations in stable isotopes of predatory fish and their representative prey fish it was revealed that the δ13c and δ15n ratios were significantly different among species and between estuaries (permanova, pseudo-f=27.3, p=0.001; table 5 and figure 4). moreover, there were significant interactions figure 3. index of dietary niche breadth (ba) of predatory fish species, arius africanus, epinephelus coioides, pomadasys argenteus and sillago sihama collected from wami and pangani estuaries. the dotted line indicates the classes of the ba, 0 – 0.39 = low ba; 0.4 – 0.6 = moderate ba and 0.61 – 1 = high ba. table 2. two-way permanova of diet composition in predatory fish species of arius africanus, epinephelus coioides, pomadasys argenteus and sillago sihama from wami and pangani estuaries. sq.root cov = square root component of variation. source of variation df ms pseudo-f p(perm) sq. root cov estuary 1 10257 4.0 0.002 11.43 fish species 3 53316 21.0 0.001 33.23 estuary x fish species 1 7927.5 3.1 0.012 23.51 residual 213 2538 50.38 table 3. pair-wise permanova of spatial variations in diet composition of arius africanus, epinephelus coioides, pomadasys argenteus and sillago sihama from wami and pangani estuaries. estuary comparisons of fish species denominator df t p a. africanus (w) and a. africanus (p) 88 1.88 0.02 e. coioides (w) and e. coioides (p) 46 2.03 0.01 p. argenteus (w) and s. sihama (p) 88 4.17 0.001 no nco mm er cia l u se on ly t.j. lyasenga et al.90 between species and estuaries (permanova, pseudof=22.6, p=0.001; table 5). at spatial scale, the individual predatory fish species exhibited a clear distinction of isotopic values between estuaries (figure 4 and table 6). as such, the mean isotope ratio of a. africanus (-17.3±1.2‰; 13.3±0.3‰) and e. coioides (-18.6±0.7‰; 13.6±0.5‰) of wami estuary were significantly lower in terms of δ13c but higher in terms of δ15n ratios when compared to that of pangani estuary (-16.3±2.1; 11.1±1.8‰; -18.0±0.7‰, 12.8±0.6‰ (figure 4 and table 6; pair-wise pertable 4. niche overlaps between pairs of predatory fish species analyzed from wami (w) and pangani (p) estuaries. species a. africanus (w) e. coioides (w) p. argenteus (w) a. africanus (p) e. coioides (p) s. sihama (p) a. africanus (w) 0.42 0.33 0.94 0.42 0.58 e. coioides (w) 0.34 0.25 0.97 0.56 p. argenteus (w) 0.37 0.84 0.21 a. africanus (p) 0.31 0.30 e. coioides (p) 0.49 s. sihama (p) table 5. two-way permanova results on isotopic variation of predatory fish species at species and estuary level for arius africanus, epinephelus coioides, pomadasys argenteus and sillago sihama from wami and pangani estuaries. source df ms pseudo-f p(perm) fish species 3 20.85 27.234 0.001 estuary 1 12.43 16.242 0.001 fish species x estuary 3 17.30 22.606 0.001 residual 95 0.766 table 6. pair-wise permanova results of intraspecific variations in stable isotopes composition of arius africanus, epinephelus coioides, pomadasys argenteus and sillago sihama between wami and pangani estuaries. estuary comparisons of fish species denominator df t p a. africanus (w) and a. africanus (p) 30 4.37 0.001 e. coioides (w) and e. coioides (p) 16 2.61 0.004 p. argenteus (w) and s. sihama (p) 23 8.58 0.001 v. buchanani (w) and v. buchanani (p) 26 1.03 0.363 figure 4. bi-plot of mean and standard error (se) of carbon and nitrogen stable isotopes of arius africanus, epinephelus coioides, pomadasys argenteus and sillago sihama and their main prey fish valamugil buchanani from wami and pangani river estuaries. no nco mm er cia l u se on ly estuarine fish trophic dynamics and anthropogenic disturbances 91 manova, t=4.4, p≤0.004). similarly, the stable isotope ratio of s. sihama based in wami when compared with p. argenteus of pangani, were significantly variable. this meant that the isotopic ratios of s. sihama of pangani estuary exhibited the lowest δ13c (-20.1±1.7‰) and higher δ15n (12.8±0.5‰) values when compared to that of p. argenteus from wami estuary (-15.5±0.2‰, 12.4±0.1‰; figure 4). for the potential prey-fish, v. buchanani despite the δ13c and δ15n mean values being slightly higher in wami compared to that of pangani (figure 4), the isotopic variations were statistically not significant (pairwise permanova, t=1.03, p>0.05; table 6). isotopic trophic diversity, niche width and overlaps of fish populations the cr (that measures the trophic diversity) and nr (measuring trophic length and diversity) of a. africanus were shorter in the wami estuary compared to that of the pangani estuary (table 7). the situation was also similar for other metrics of trophic niche width of a. africaus, such as the bayesian estimate of trophic niche width (seab) and seac (figure 5 and table 7). likewise, the same trend was also noticed for the seab and seac of the other three pairs of species of comparison from wami and pangani estuaries (figure 5 and table 7). the results of the intraand interspecific comparisons of seab within and between the estuaries varied from one pair to another of predatory species. the seab of a. africanus from wami was significantly smaller than that of pangani estuary with no intraspecific overlap in between bayesian estimate of 95% ci (table 7). this was also in line with estimate of the probability of seab of less than 1% for the a. africanus based in pangani to be smaller than that of wami estuary (table 7). the seab of e. coioides from both wami and pangani estuaries were statistically similar to each other due to large overlaps between 95% ci and 50% in magnitude of probability of the seab for the two pairing species (table 7). even cr and nr of e. coioides table 7. bayesian standard ellipse area widths (seab), magnitude of probability of seab for the wami-based species to be smaller than pangani based estuarine species; seac, cr and nr for a. africanus, e. coioides, s. sihama, p. argenteus and v. buchanani studied in wami and pangani estuaries. ci – credible interval of 95%. species wami pangani probability (%) of seab for species of wami < pangani for 95%ci %ci lower upper %ci lower upper a. africanus 99 0.59 2.31 99 1.99 8.04 95 0.68 1.92 95 2.36 6.67 0.0005 50 0.95 1.35 50 3.20 4.60 mode 1.11 3.81 seac 1.26 3.69 cr 3.56 7.00 nr 0.92 3.44 e. coioides 99 0.37 3.09 99 0.44 2.86 95 0.47 2.27 95 0.56 2.20 50 0.78 1.29 50 0.89 1.38 mode 1.00 1.09 seac 1.27 1.32 cr 1.86 1.51 nr 1.50 1.46 p. argenteus (wami)/s. sihama (pangani) 99 0.02 0.11 99 0.84 4.03 95 0.02 0.09 95 0.99 3.20 50 0.04 0.05 50 1.49 2.18 50 mode 0.04 1.81 seac 0.05 1.98 cr 0.61 4.38 nr 0.26 1.52 v. buchanani 99 0.87 4.36 99 1.34 5.89 95 1.06 3.53 95 1.67 4.77 50 1.57 2.35 50 2.39 3.37 mode 1.91 2.85 seac 2.18 3.19 cr 2.4 2.98 nr 2.55 2.56 no nco mm er cia l u se on ly t.j. lyasenga et al.92 from these estuaries were relatively similar. the seab, and thus, trophic niche width, of s. sihama of pangani was significantly larger as compared to that of a. argenteus of wami due to zero overlap between the 95% ci (table 7). contrary to that, the substantial overlap between the 95% ci for the v. buchanani from both wami and pangani estuary (table 7) was considered as an indication that the trophic niche width of the species in these habitats is somewhat similar. furthermore, the seab between the following pairs of species were also significantly different when their 95% cis were compared: both pairs of a. africanus versus e. coioides from within pangani and wami estuaries, e. coioides versus p. argenteus of wami as well as a. africanus versus p. argenteus of wami estuary (table 7). similar to a. africanus, the cr and nr of s. sihama (cr=4.4‰, nr=1.5‰) and v. buchanani (cr=3.0‰, nr=2.6‰) were consistently higher in pangani estuary than those of their comparable species in the wami estuary (p. argenteus cr=0.6‰, nr=0.3‰; and v. buchanani cr=2.4‰, nr=2.5‰; figure 5 and table 7). the seac showed substantial overlap between the standard ellipses for a. africanus and e. coioides from wami estuary, and between all comparable species from pangani estuary (figure 5). this indicated that the degree of dietary resource sharing by predatory fish was higher in pangani compared to that of wami estuaries. specifically, the estimates of seac overlap analyses revealed that, in wami estuary, the percentage of overlap between seac for a. africanus onto that of e, coioides, and vice versa, was 54%. no overlap of seac was found between a. africanus versus p. argenteus nor e. coioides versus p. argenteus from wami estuary. contrary to that, in pangani estuary, the largest percentages of overlap were indicated by seac for the s. sihama onto that of a. africanus (62%) and for the e. coioides onto that of a. africanus (61%). the percentage of seac for the a. africanus versus e. coiodes was 21% and a. africanus versus s. sihama was 34%. more so, the percentage of e. coioides that overlapped with s. sihama was 58%, while that of s. sihama onto the seac of e. coioides was 16%. the results of intra-specific overlap between estuaries indicated that the seac for the a. africanus of pangani onto that of wami was 11%, but the values of the overlap for the vice versa was 33%. for the e. coioides, the overlap percentages between estuaries were 57% and 59%. again, there was no ellipse overlap for the p. argenteus of wami and s. sihama of pangani. trophic niche and redundancy at estuary level the seac and ta that infer the total trophic niche area of fish community were smaller in wami (seac=5.4; ta=14.6) contrary to that of pangani estuaries (seac=7.6; ta=25.1; figure 6). the comparison test for the bayesian estimate 95% ci of the two estuaries revealed that seab of wami was slightly smaller than that of pangani as they showed substantial overlap between the 95% ci (wami estuary 3.90-6.94 of 95% ci; pangani estuary 5.70-9.69 of 95% ci). however, the magnitude of probability for seab of pangani to be smaller than that of wami was impossible to occur (<1%). more so, the high level of overlap between seac of two estuaries was noticed whereby the seac of wami was greatly enclosed within that of pangani estuary (figure 6). as well, the layman community indices showed that the two estuaries differ in trophic resources and level of trophic diversifications of the same fish species investigated (figure 7). the trophic length as indicated by the nr of pangani estuary (nr=2.45) was marginally shorter when correlated with that of wami (nr=2.91; figure 7). the opposite situation was noticed for the indicator of figure 5. the convex hull area (dotted lines) and standard ellipse areas (sea) measuring the trophic niche width of individual predatory fish: 1.1 arius africanus from pangani; 2.5 a. africanus of wami; 1.2 epinephelus coioides of pangani; 2.6 e. coioides of wami; 1.3 sillago sihama of pangani; 2.7 pomadasys argenteus of wami and their prey fish; 1.4 = valamugil buchanani from pangani and 2.8 = v. buchanani of wami and pangani estuaries. no nco mm er cia l u se on ly estuarine fish trophic dynamics and anthropogenic disturbances 93 basal and average trophic diversity of fish as expressed by the cr and cd (figure 7). the trophic redundancy (measured as m-nnd) and its standard deviation (sdnnd) differed substantially between the two estuaries with m-nnd being smaller and sdnnd larger in wami than pangani estuary (figure 7). discussion the findings of this study indicate how anthropogenic activities in the upstream of the river catchments induce changes to the estuarine ecosystem’s functioning. the results suggest that despite the wami estuary being under conservation status, the high level of disturbances upstream of the river increase vulnerability of the estuarine food web structure and ecosystems functioning. this is probably connected to a lower primary production, to changes in trophic positions of consumers, and thus to modifications occurring in the trophic niche of the predatory fish. as a result, the reduced trophic diversity, ecological redundancy, and complexity of the ecosystem are most likely experienced in wami estuary than pangani estuary. this was different from the expectation as the considered highly disturbed pangani estuary presented a relatively stable, resilient, and ecologically redundant ecosystem. even though our findings are based on few components of the estuarine food webs, they include important compartments that greatly shape the complexity and thus determine the ecological redundancy of the estuarine ecosystems of concern. furthermore, the inferences drawn from the present study are based on data collected in a specific seasonal period, so that isotopic values gathered from estuarine fish with large body size (i.e., predatory fish) mirror the exploitation of natural resources during their growth period. moreover, data collection that adheres to seasons may elucidate more on long-term changes in the use of dietary resources and other niche characteristics of estuarine fish assemblage. moreover, our results explain how both predatory and prey fish respond to the dynamics of dietary resources, representing with a good approximation the food web state in tropical estuaries differing in the degree of anthropogenic disturbances. our findings showed that both stomach contents and stable isotopes can reveal significant differences in diet and isotopic values among fish species in different estuaries. matching of these findings confirm that both methods can characterize the foraging flexibility of predatory fish species of marine origin in estuaries. except for a. africanus, the results of the index of dietary niche breadth, ba, have matched well with that of isotopic trophic niche metrics of the species between estuaries. as such, the observed consistent highest ba of e. coioides in both estuaries correlated with the similarity in their seab between figure 6. trophic niche width indicated by ta (dotted lines) and sea (solid lines) for the pooled stable isotopes of carbon and nitrogen of fish from wami and pangani estuaries (n = 103). cycle points represent individuals of all the species measured in each estuary and standard ellipse/circle represent the maximum likelihood of isotopic niche of the species analyzed. figure 7. the estuarine community-wide layman metrics of wami and pangani estuaries that showed the (i) trophic length, or complexity, nr (dy_range), (ii) trophic diversity, or cr (dx_range), (iii) niche width, or convex hull area (ta), (iv) average trophic diversity, or mean distance to the centroid (cd), (v) trophic redundancy, or the mean nearest neighbor distance (nnd), and (vi) evenness of species packing, or extent of trophic diversification, that is, the standard deviation ) of ndd (sdnnd). black dots are the modes, boxes indicate the credible intervals at 50%, 75% and 95%. the numbers above each box are modes; the cross (red) is the values of the true population. no nco mm er cia l u se on ly t.j. lyasenga et al.94 estuaries. similarly, the estimated moderate ba of s. sihama from pangani and the lowest ba of p. argenteus from wami correlated with the larger seab of panganibased s. sihama versus the comparable smaller seab of p. argenteus of wami estuary. however, the mismatch of trophic feeding niche of a. africanus by the two methods indicates that the trophic dynamics in highly fluctuating estuarine ecosystems requires robust methods that should complement the conventional stomach content analysis for drawing a sound and logical conclusion. the application of stable isotope in trophic ecology studies is widely accepted (e.g. pasquaud et al., 2010; olsen et al., 2011; layman et al., 2012; cummings et al., 2012) as it is among the powerful tools that ensure the determination of the integrated diet consumed and assimilated at a coarse taxonomic scale but over a long period (michener and lajtha, 2007; selleslagh et al., 2015). relative to that, the stomach content analysis determines the diet ingested instantly by the consumers at a finer taxonomic level (mahesh et al., 2018). the marked variations in the dietary composition and trophic niche among the predatory fish species found in our study imply that they largely consume similar preys but in different proportions in the two estuaries. differences in the amounts of preys consumed may be linked to the prey abundance and catchability along with prey visibility by predatory species, plus the feeding strategies and energy requirements (kulbicki et al., 2005; kroetz et al., 2016). both stomach contents and stable isotope-derived layman metrics showed a considerable degree of trophic resource sharing and thus trophic niche overlaps among these predatory fish in estuaries. this finding agrees with those from the study of matich et al. (2017). this possibly does not imply competition in trophic resources but rather the prey-catch strategies that define the prey types consumed. for instance, the higher trophic niche breadth and similarity in isotopic niche width of e. coioides in wami and pangani estuaries could be favored by its ambush strategy and structural complexity of the estuaries (gibran, 2007). complex habitats such as these mangrove-sheltered estuaries provide enhanced cover that reduces the possibilities of predators to be detected by the preys (mwijage et al., 2018a). this situation is probably connected with the highest overlap of seac of the species between estuaries. interand intra-species variations in isotopic values could mean considerable individual variability in the feeding strategies with individuals consuming prey of different trophic spectra and the use of basal nutritional sources they rely on. differences in both dietary niche breadth and isotopic niche width at the estuary level suggest that fish species may have high trophic niche plasticity or flexibility, which is indicated by matich et al. (2017) as a mechanism to respond to natural and human-induced environmental change. this is further indicated by the dissimilarity in magnitude of the isotopic metric-based overlap of each possible pair of the predatory fish species in the two estuaries examined. the highest degree of overlap for the fish species of pangani might be an indication of shifts in the trophic positions of organisms preyed upon by these species in two estuaries resulting to the observed differences in trophic length, diversity and trophic niches justified by nr, cr, cd and seac/seab. the shifts in trophic positions of consumers within comparable aquatic environment is mainly driven by differences in individual diet, trophic discrimination or enrichment factor (villamarín et al., 2018), as well as differences in the level of nutrients load fueling the base of the food web (warry et al., 2016). furthermore, our findings suggest a possibility of presence of ontogenetic shift in trophic positions (park et al., 2018; villamarín et al., 2018) for the comparable estuaries examined herein. although the length size of the individuals was not considered during analysis, it might also influence the results of the differences in isotopic values of the species between the estuaries. this is mainly linked with the size range (total length, tl) of a. africanus from wami versus of pangani and p. argenteus of wami versus s. sihama of pangani. the literatures emphasize that resource utilization patterns of fish change markedly with ontogeny (davis et al., 2012; park et al., 2018; villamarín et al., 2018). specifically, with the exclusion of e. coioides, the larger cr of predatory fish populations of pangani estuary might be linked with their prey to feed on the diet with varied nutritional sources. the largest seab of a. africanus population of pangani could be probably associated with a high level of opportunistic or generalist feeding mechanism as a coping strategy resulting from exposing the estuary to more frequent human perturbations. the smallest seab of p. argenteus of wami and its δ13c mean values (-15.5±0.2‰) is an indication that the riverine or terrestrial derived dietary sources of which their δ13c ratio range from -23 to -30‰ (bouillon et al., 2011) had little influence on the diet dynamics of this species. this is somewhat contributed by the sampling location of the samples. pomadasys argenteus was caught in the estuarine plume away from the river mouth, and this environment might have been less influenced by the riverine dietary sources. the same situation further explains the observed non-overlap of seac between p. argenteus and any other pair of predatory fish studied. more so, the similarity in seab of e. coioides between estuaries suggests low trophic niche flexibility of the species. in view of that situation, the environmental filtering hypothesis plays a role in this assembly that was not similar due to divergence in the environmental conditions of the species habitats (pereira et al., 2017). no nco mm er cia l u se on ly estuarine fish trophic dynamics and anthropogenic disturbances 95 furthermore, the higher values of δ15n and nr of wami-associated species may be an indication of the predatory fish to occupy higher trophic position usually connected to the nature of the diet assimilated by individuals, the trophic enrichment factor and high levels of nutrients load (villamarín et al., 2018), as opposed to pangani estuary. the effects of high values of δ15n at low trophic level are transmitted upwards higher up to top predators of estuarine food web, thus affecting the overall trophic organization (warry et al., 2016). parallel to this, the substantial low δ13c ratio could be contributed by the wami estuarine food web to be fed by riverine primary food sources and high inorganic nitrogen load that elevate δ15n values in secondary production (michener and lajtha, 2007; woodland and secor, 2011; warry et al., 2016). in the pangani estuary, water abstractions upstream of the river due to multi-reservoirs for hydropower generation and irrigation is linked with trophic interactions downstream the estuarine and coastal food webs (mwijage et al., 2018b). these human activities contributed to modify the nutrients’ biogeochemistry in the estuary and increased retention time of sediment, organic particles and accompanied nutrients in the reservoirs, upstream of pangani river (selemani et al., 2017). eventually, pangani estuary experience lower δ13c values of terrestrial derived basal food sources (mwijage et al., 2018a) and regulated nutrients load including nitrogen compounds. at the estuarine level, the differences in mean distance to the centroid (cd) and seac of the wami and pangani fish species may be linked to the variations in estuarine primary productivity. the high loads of suspended sediments in the wami estuary as reported by kiwango et al. (2015) contribute to reducing trophic diversity and abundance resulting from the low rate of aquatic primary production (selemani et al., 2017). light penetration is constrained by the suspended particles that, consequently, limit photosynthetic activities. in that way, trophic web connectance, or the number of trophic links to the primary producers, tend to be small. this also may be a factor contributing to the low level of trophic overlap of wami and consequently, low trophic redundancy (de carvalho et al., 2017; lira et al., 2018). the overall effect is to make food web components vulnerable to secondary extinctions as a result of the reduced complexity of the ecosystem (lira et al., 2018). this supports our findings but not what was expected of wami to exhibit higher trophic redundancy than pangani. as such, wami estuary presented a large sd-nnd which suggests a reduced trophic redundancy. this is in contrast to the pangani estuary that according to hellar-kihampa et al. (2013), have relatively low sediment loads as large volumes are retained in the dams located upstream of the estuary. this concurred with the findings of abrantes et al. (2014) who reported that the high sediment loads affect aquatic primary productivity and hence small seac of fish populations and communities in the betsiboka and tana estuaries. the predatory fish assemblages in the wami estuary are probably less opportunistic in feeding that implies a low level of resilience upon disturbances. selleslagh and amara (2014) also showed that fish in the canche estuary showing very low anthropogenic pressure, exhibited a specialist feeding strategy. the resilience of the estuarine ecosystem is contributed by among others, trophic complexity or multiplicity of trophic linkages along the trophic web and species diversity (lira et al., 2018). in that sense, the present study agreed with the findings of pasquaud et al. (2010) and donázar-aramendía et al. (2019) who showed that high degrees of anthropogenic pressures modify the structural components and ecological roles of the species in estuarine ecosystems. structural modifications in wami estuary are certainly influenced by the high turbidity that, apart from affecting primary production, hampers species of higher-order consumers or predatory fish because of lower performances in prey hunting. the reduction of predatory fish species from a community lead to the dominance of its prey at the lower trophic level and subsequent secondary extinctions of other carnivores (sanders et al., 2018). this also leads to a reduction in trophic redundancy and consequently complexity and stability of the estuarine food webs. on the other hand, the large degree of overlaps in the isotopic niche of species from the pangani estuary could not imply the existence of potential competitions of the trophic resources but rather the similarity in environmental conditions in microhabitats the fish species used for feeding. greater diet overlaps increase trophic redundancy of the ecosystem (sanders et al., 2018), a contention that seems to agree with our results which show that the pangani estuary has greater trophic redundancy than that of wami. conclusions overall, the community-wide metrics derived from stable isotopes when complemented with dietary data elucidate how predatory fish and their prey respond to human-induced changes in trophic niche width; this response determines cascading effects that impact the trophic redundancy and complexity of the estuarine ecosystems. our findings indicate that anthropogenic activities in the upstream modified higher trophic position of fish consumers in wami relative to that of pangani estuaries. thus, without management interventions, food web structure reorganization and disruption are likely to occur within the above-mentioned estuaries. moreover, our findings should be taken with the caveat that not all components of the estuarine food webs were used to draw such inferences, but they could be used as indicators of showing that the two estuarine systems investigated are ecologically disturbed no nco mm er cia l u se on ly t.j. lyasenga et al.96 by anthropogenic activities. therefore, management initiatives should be strengthened for maintaining the structures and functioning of these fragile ecosystems. corresponding author: alistidia paul mwijage, tanzania fisheries research institute (tafiri) – headquarters, p. o. box 9750, dar es salaam, tanzania. e-mail: alistidiamwijage@tafiri.go.tz; zawadi2007@gmail.com. key words: diet; predatory fish; stable isotope; trophic niche; layman community metrics; anthropogenic disturbances funding: the university of dar es salaam through the coafaqf19058 project number covered all costs from the study design through data collection and interpretation of results. acknowledgements: the authors acknowledge the funder of this study, the university of dar es salaam that covered all costs from the study design through data collection and interpretation of the results. the authors also recognize and appreciate the technical support and research facilities provided by the chemistry department of the college of natural and applied sciences and school of aquatic sciences and fisheries technology of the university of dar es salaam. the support of dr. blandinga lugendo (from university of dar es salaam) who assisted the logistics of identification of a laboratory in united kingdom to undertake the stable isotope analysis is also appreciated. the authors also acknowledge and appreciate the technical support and research facilities provided oea labs limited in the united kingdom. authors’ contributions: all authors played different roles throughout the production of this manuscript. project conceptualization was conducted by tjl, das, jamm and lg. data curation was conducted by plm and das. methodology and formal data analysis were conducted by plm, apm, tjl and das. manuscript writing (original draft) was done by apm, tjl, das and pjm. visualization and the first draft manuscript review were performed by das, jamm, lg and das. funding was collectively solicited by plm, das, jamm, lg and das. conflict of interest: the authors declare no potential conflict of interest. availability of data and materials: all data generated or analyzed during this study are included in this published article. received: 18 july 2021. accepted: 24 november 2021. this work is licensed under a creative commons attribution noncommercial 4.0 license (cc by-nc 4.0). ©copyright: the author(s), 2021 licensee pagepress, italy advances in oceanography and limnology, 2021; 12:9987 doi: 10.4081/aiol.2021.9987 references abrantes kg, barnett a, bouillon s, 2014. stable isotope-based community metrics as a tool to identify patterns in food web structure in east african estuaries. funct. ecol. 28:270-282. abrantes kg, barnett a, marwick tr, bouillon s, 2013. importance of terrestrial subsidies for estuarine food webs in contrasting east african catchments. ecosphere. 4:14. anderson mj, gorley rn, clarke kr, 2008. permanova+ for primer: guide to software and statistical methods primer-e. anderson ep, mcnally c, kalangahe b, gutmann roberts c, britton jr, 2018. quantifying trophic interactions and niche sizes of juvenile fishes in an invaded riverine cyprinid fish community. ecol. freshw. fish 27:976-987. anderson ep, mcnally c, kalangahe b, ramadhani h, mhitu h, 2007. a rapid ecological assessment of the wami river estuary, tanzania. available at: https://pdf.usaid.gov/pdf_ docs/pnadn793.pdf : 38 pp. baeta a, valiela i, rossi f, pinto r, richard p, niquil n, et al., 2009. eutrophication and trophic structure in response to the presence of the eelgrass zostera noltii. mar. biol. 156:21072120. beck mw, heck kl, able kw, childers dl, eggleston db, 2001. the identification, conservation and management of estuarine and marine nurseries for fish and invertebrates. bioscience 51:633-641. bianchi g, 1995. fao species identification sheets for fishery purposes. field guide to the commercial marine and brackish-water species of tanzania. prepared and published with the support of tcp/urt/4406 and fao (firm) regular programme. rome, italy. bouillon s, connolly rm, and gillikin dp, 2011. use of stable isotopes to understand food webs and ecosystem functioning in estuaries. in: wolanski e, mclusky ds (eds.), treatise on estuarine and coastal science, 7:143-173. carbonara p, follesa mc, (eds.) 2019. handbook on fish age determination: a mediterranean experience. studies and reviews. fao. no. 98. rome, italy : 192 pp. cole ml, kroeger kd, mcclelland jw, 2016. effects of watershed land use on nitrogen concentrations and δ15 nitrogen in groundwater. biogeochemistry. 77:199-215. cummings do, buhl j, lee rw, simpson sj, holmes sp, 2012. estimating niche width using stable isotopes in the face of habitat variability: a modelling case study in the marine environment. plos one. 7:e40539. davis am, blanchette ml, pusey bj, jardine td, pearson rg, 2012. gut content and stable isotope analyses provide complementary understanding of ontogenetic dietary shifts and trophic relationships among fishes in a tropical river. freshw. biol. 57:2156-2172. de carvalho dr, de castro dmp, callisto m, moreira mz, pompeu ps, 2017. the trophic structure of fish communities from streams in the brazilian cerrado under different land uses: an approach using stable isotopes. hydrobiologia 795:199-217. dierking bj, meyer al, 2009. short communication: prey regurgitation in the grouper cephalopholis argus. j. of appl. ichthy. 25:600-602. donázar-aramendía i, sánchez-moyano je, garcía-asencio i, miró jm, megina c, garcía-gómez jc, 2019. human pressures on two estuaries of the iberian peninsula are reflected in food web structure. sci rep. 9:1-10. eeden a, mehta l, koppen b, 2016. whose waters? large-scale agricultural development and water grabbing in the wamiruvu river basin, tanzania. water altern. 9:608-626. no nco mm er cia l u se on ly estuarine fish trophic dynamics and anthropogenic disturbances 97 elliott m, whitfield ak, potter ic, blaber sjm, cyrus dp, nordlie fg, et al., 2007. the guild approach to categorizing estuarine fish assemblages: a global review. fish fish. 8:241-268. euzen o. 1987. food habits and diet composition of some fish of kuwait. bull. mar. sci. 9:65-85. fiu-glows, 2016. freshwater inflow requirements for wami river estuary, saadani national park, tanzania. : 104 pp. garcía k, sanpera c, jover ll, palazón s, gosálbez j, górski k, et al., 2020. invasive mink does not drive trophic displacement of the native mink during an invasion process. animals. 10: 1387. gibran fz, 2007. activity, habitat use, feeding behavior, and diet of four sympatric species of serranidae (actinopterygii: perciformes) in southeastern brazil. neotrop. ichthyol. 5:387-398. gillson j, 2011. freshwater flow and fisheries production in estuarine and coastal systems: where a drop of rain is not lost. rev. fish sci. 19:168-186. gutmann roberts c, britton jr 2018. quantifying trophic interactions and niche sizes of juvenile fishes in an invaded riverine cyprinid fish community. ecol. freshw. fish 27: 976-987. han r, chen q, wang l, tang x, 2016. preliminary investigation on the changes in trophic structure and energy flow in the yangtze estuary and adjacent coastal ecosystem due to the three gorges reservoir. ecol inform. 36:152-161. harrison td, whitfield ak, 2006. temperature and salinity as primary determinants influencing the biogeography of fishes in south african estuaries. estuar. coast shelf sci. 66: 335345. hellar-kihampa h, de wael k, lugwisha e, van grieken r, 2013. water quality assessment in the pangani river basin, tanzania: natural and anthropogenic influences on the concentrations of nutrients and inorganic ions. int. j. river basin manag. 11:55-75. hoeinghaus dj, vieira jp, costa cs, bemvenuti ce, winemiller ko, garcia am, 2011. estuary hydrogeomorphology affects carbon sources supporting aquatic consumers within and among ecological guilds. hydrobiologia. 673:79-92. hurlbert sh, 1978. the measurement of niche overlap and some relatives. ecology 59:67-77. hyslop ej, 1980. stomach contents analysis-a review of methods and their application. j. fish biol. 17:411-429. jackson al, inger r, parnell ac, bearhop s, 2011. comparing isotopic niche width among and within communities: siber – stable isotope bayesian ellipses in r. j. animal ecol. 80:595-602. jackson al, parnell ac, inger r, bearhop s, 2011. comparing isotopic niche widths among and within communities: siber stable isotope bayesian ellipses in r. j. anim. ecol. 80:595-602. kiwango h, njau kn, wolanski e, 2015. the need to enforce minimum environmental flow requirements in tanzania to preserve estuaries: case study of mangrove-fringed wami river estuary. ecohydrol. hydrobiol. 15:171-181. kiwango h, njau kn, wolanski e, 2018. the application of nutrient budget models to determine the ecosystem health of the wami estuary, tanzania. ecohydrol. hydrobiol. 18:107119. kostecki c, le loc’h f, roussel jm, desroy n, huteau d, riera p, et al., 2010. dynamics of an estuarine nursery ground: the spatio-temporal relationship between the river flow and the food web of the juvenile common sole (solea solea, l.) as revealed by stable isotopes analysis. j. sea res. 64:54-60. kroetz a, drymon jm, powers sp, 2016. comparative dietary diversity and trophic ecology of two estuarine mesopredators. estuaries and coasts. 40:1171-1182. kulbicki m, bozec y, labrosse p, letourneur y, mou-tham g, wantiez l, 2005. diet composition of carnivorous fishes from coral reef lagoons of new caledonia. aquat. living resour. 18:231-250. layman ca, araujo ms, boucek r, hammerschlag-peyer cm, harrison e, jud zr, et al., 2011. applying stable isotopes to examine food-web structure : an overview of analytical tools. biol. rev. 87:545-562. layman ca, arrington da, montana cg, post dm, 2007. can stable isotope ratios provide for community-wide measures of trophic structure? ecology 88:42-48. lira a, angelini r, le loc’h f, ménard f, lacerda c, frédou t, et al., 2018. trophic flow structure of a neotropical estuary in north-eastern brazil and the comparison of ecosystem model indicators of estuaries. j. mar. syst. 182:31-45. mancinelli g, vizzini s, 2015. assessing anthropogenic pressures on coastal marine ecosystems using stable cns isotopes: state of the art, knowledge gaps, and community-scale perspectives. estuar. coast shelf sci. 156:195-204. matich p, kiszka jj, gastrich kr, heithaus mr, 2017. trophic redundancy among fishes in an east african nearshore seagrass community inferred from stable-isotope analysis. j. fish biol. 91:490-509. michener r, lajtha k (eds.), 2007. stable isotopes in ecology and environmental science. blackwell publishing ltd. : 561 pp. mwijage ap, shilla da, machiwa jf, 2018. differences in trophic resources and niches of two juvenile predatory species in three pangani estuarine zones, tanzania: stomach contents and stable isotope approaches. j. biol res. 25:13. mwijage ap, shilla da, machiwa jf, zhang j, fuqiang w, ying w, 2018. important organic matter sources and trophic pathways for the nutrition of hilsa kelee (cuvier, 1829) and valamugil buchanani (bleeker, 1853) in pangani macrotidal estuary, tanzania. chem ecol. 34:941-963. olsen ys, fox se, teichberg m, otter m, valiela i, 2011. 15n and δ13c reveal differences in carbon flow through estuarine benthic food webs in response to the relative availability of macroalgae and eelgrass. mar. ecol. prog. ser. 421:83-96. park hj, park th, lee c il, kang ck, 2018. ontogenetic shifts in diet and trophic position of walleye pollock, theragra chalcogramma, in the western east sea (japan sea) revealed by stable isotope and stomach content analyses. fish. res., 204:297-304. pasquaud s, david v, lobry j, girardin m, sautour b, elie p, 2010. exploitation of trophic resources by fish under stressful estuarine conditions. mar. ecol. prog. ser. 400:207-219. pereira ls, mise ft, tencatt lfc, baumgartner mt, agostinho aa, 2017. is coexistence between non-native and native erythrinidae species mediated by niche differentiation or environmental filtering? a case study in the upper paraná river floodplain. neotrop. ichthyol. 15:1-12. no nco mm er cia l u se on ly t.j. lyasenga et al.98 peterson bj, fry b, 1987. stable isotopes in ecosystem studies. annu. rev. ecol. syst. 18:293-320. pihl l, codling i, mathieson s, mclusky ds, roberts c, 2002. habitat use by fishes in estuaries and other brackish areas, p.10-53. in: m. elliott, k. hemingway (eds.), fishes in estuaries. blackwell publishing ltd. post d, layman c, arrington d, takimoto g, quattrochi j, montana c, 2007. getting to the fat of the matter: models, methods and assumptions for dealing with lipids in stable isotope analyses. oecologia. 152:179-89. post dm, 2002. using stable isotopes to estimate trophic position: models, methods, and assumptions. ecology. 83:703718. pwbo/iucn, 2008. fish and invertebrate life histories and important fisheries of the pangani river basin. pangani river basin flow assessment. pangani basin water board, moshi and iucn eastern and southern africa regional programme: 63 pp. ramadhani h, mhitu h, 2007. a rapid ecological assessment of the wami river estuary, tanzania: 16 pp. rooney n, mccann ks, 2012. integrating food web diversity, structure and stability. trends ecol. evol. 27:40-46. rossi l, lascio a, carlino p, calizza e, costantini ml, 2015. predator and detritivore niche width helps to explain biocomplexity of experimental detritus-based food webs in four aquatic and terrestrial ecosystems. ecol. complex. 23:1424. sanders d, thébault e, kehoe r, frank van veen fj, 2018. trophic redundancy reduces vulnerability to extinction cascades. pnas 115:2419-2424. selemani jr, zhang j, muzuka ann, njau kn, zhang g, mzuza mk, et al., 2017. nutrients’ distribution and their impact on pangani river basin’s ecosystem tanzania. environ. technol. 0:1-15. selleslagh j, amara r, 2014. are estuarine fish opportunistic feeders? the case of a low anthropized nursery ground (the canche estuary, france). estuar. coasts. 38:252-267. selleslagh j, blanchet h, bachelet g, lobry j, 2015. feeding habitats, connectivity and origin of organic matter supporting fish populations in an estuary with a reduced intertidal area assessed by stable isotope analysis. estuar. coasts. 38:1431-1447. semba m, kimirei i, kyewalyanga m, peter n, brendonck l, somers b, 2016. the decline in phytoplankton biomass and prawn catches in the rufiji-mafia channel, tanzania. western indian ocean j. mar. sci. 15:15-29. shaghude yw, 2016. estuarine environmental and socio-economic impacts associated with upland agricultural irrigation and hydropower developments: the case of rufiji and pangani estuaries, tanzania. in: s. diop, p. scheren, jf. machiwa (eds.), estuaries: a lifeline of ecosystem services in the western indian ocean: 13 pp. shaw al, frazier bs, kucklick jr, sancho g, 2016. trophic ecology of a predatory community in a shallow-water, highsalinity estuary assessed by stable isotope analysis. mar. coast fish. 8:46-61. vander-zanden mj, rasmussen jb, 2001. variation in δ15n and δ13c trophic fractionation: implications for aquatic food web studies. limnol. oceanogr. 46:2061-2066. vignon m, dierking j. 2011. prey regurgitation and stomach vacuity among groupers and snappers. environ. biol. fishes 90:361-366. villamarín f, jardine td, bunn se, marioni b, magnusson we, 2018. body size is more important than diet in determining stable-isotope estimates of trophic position in crocodilians. sci. rep. 8:1-11. wang j, chapman xu, wang jy, gu b, 2018. isotope niche dimension and trophic overlap between bigheaded carps and native filter-feeding fish in the lower missouri river, usa. plos one, 13:1-13. warry fy, reich p, cook plm, mac nally r, thomson jr, woodland rj, 2016. nitrogen loads influence trophic organization of estuarine fish assemblages. funct. ecol. 30:17231733. whitfield ak, 2016. biomass and productivity of fishes in estuaries: a south african case study. j. fish biol. 89:19171930. woodland rj, secor dh, 2011. differences in juvenile trophic niche for two coastal fish species that use marine and estuarine nursery habitats. mar. ecol. prog. ser. 439:241-254. zhang m, wang y, gu b, li y, zhu w, zhang l, et al., 2019. resources utilization and trophic niche between silver carp and bighead carp in two mesotrophic deep reservoirs. j. freshw. ecol. 34:199-212.no nco mm er cia l u se on ly layout 1 introduction concerns regarding the presence of cyanobacterial toxins (cyanotoxins) in drinking water and associated health effects have raised research and public health interest worldwide. microcystins (mcs) are probably the most frequently found cyanotoxins which can be produced by various cyanobacterial genera including water bloomand scum-forming planktonic cyanobacteria such as dolichospermum (formerly anabaena), microcystis or planktothrix (manganelli et al., 2012). cyanobacteria representing these genera have been previosly identified in ghanaian water reservoirs along with other cyanobacterial species potentially producing mcs (addico et al., 2006, 2009, 2017). mcs are highly toxic for mammals with acute ld50 as low as 50-60 µg kg–1, mouse, i.p. (bláha et al., 2009; van apeldoorn et al., 2007). their acute effetcs are primarily manifested in liver but mcs have been shown to induce gastrointestinal and renal damage or neurological symptoms as well (manganelli et al., 2012). chronic exposures to mcs have been linked to tumor promoting and carcinogenic effects which is based on laboratory animal and in vitro experiments (svircev et al., 2010) and supported also by results of epidemiologic studies of human population consuming drinking water contaminated by these cyanotoxins (fleming et al., 2002; svircev et al., 2009, 2013, 2014; ueno et al., 1996; yu et al., 1995; zhou et al., 2002). in fact, mcs have been classified as possible human carcinogen (class 2b) by the international agency for research on cancer (grosse et al., advances in oceanography and limnology, 2017; 8(1): 92-106 article doi: 10.4081/aiol.2017.6323 this work is licensed under a creative commons attribution-noncommercial 4.0 international license (cc by-nc 4.0). cyanobacteria and microcystin contamination in untreated and treated drinking water in ghana gloria naa dzama addico,1* jörg d. hardege,2 jiří kohoutek,3 k.a.a. degraft-johnson,1 pavel babica3,4 1csir water research institute, achimota, accra, ghana; 2biological science department, university of hull, united kingdom; 3recetox research centre for toxic compounds in the environment, faculty of science, masaryk university, brno, czech republic; 4department of experimental phycology and ecotoxicology, institute of botany, czech academy of sciences, brno, czech republic *corresponding author: naadzama443@hotmail.com abstract although cyanobacterial blooms and cyanotoxins represent a worldwide-occurring phenomenon, there are large differences among different countries in cyanotoxin-related human health risk assessment, management practices and policies. while national standards, guideline values and detailed regulatory frameworks for effective management of cyanotoxin risks have been implemented in many industrialized countries, the extent of cyanobacteria occurrence and cyanotoxin contamination in certain geographical regions is underreported and not very well understood. such regions include major parts of tropical west and central africa, a region constisting of more than 25 countries occupying an area of 12 million km2, with a total population of 500 milion people. only few studies focusing on cyanotoxin occurrence in this region have been published so far, and reports dealing specifically with cyanotoxin contamination in drinking water are extremely scarce. in this study, we report seasonal data on cyanobacteria and microcystin (mc) contamination in drinking water reservoirs and adjacent treatment plants located in ghana, west africa. during january-june 2005, concentrations of mcs were monitored in four treatment plants supplying drinking water to major metropolitan areas in ghana: the treatment plants barekese and owabi, which serve kumasi metropolitan area, and the plants kpong and weija, providing water for accra-tema metropolitan area. hplc analyses showed that 65% samples of raw water at the intake of the treatment plants contained intracellular mcs (maximal detected concentration was 8.73 µg l–1), whereas dissolved toxins were detected in 33% of the samples. significant reduction of cyanobacterial cell counts and mc concentrations was achieved during the entire monitoring period by the applied conventional water treatment methods (alum flocculation, sedimentation, rapid sand filtration and chlorination), and mc concentration in the final treated water never exceeded 1 µg l–1 (who guideline limit for mc-lr in drinking water). however, cyanobacterial cells (93-3,055 cell ml–1) were frequently found in the final treated water and intracellular mcs were detected in 17% of the samples (maximal concentration 0.61 µg l–1), while dissolved mcs were present in 14% of the final treated water samples (maximal concentration 0.81 µg l–1). it indicates a borderline efficiency of the water treatment, thus mc concentrations in drinking water might exceed the who guideline limit if the treatment efficiency gets compromised. in addition, mc concentrations found in the raw water might represent significant human health risks for people living in areas with only a limited access to the treated or underground drinking water. key words: cyanobacteria; cyanotoxins; drinking water treatment; microcystins; water blooms. received: october 2016. accepted: may 2017.no nco mm er cia l u se on ly microcystins in drinking water in ghana 93 2006). in addition, other epidemiological studies associated exposures to toxic cyanobacterial blooms and mcs with chronic liver damage (chen et al., 2009; li et al., 2011; zhang et al., 2015), and mcs were also implicated in neurotoxicity and neurodegenerative diseases (feurstein et al., 2009, 2010). mcs are therefore regarded as human health hazard. exposure of human beings to mcs can occur via different routes, such as recreational and sport activities in contaminated water, consumption of contaminated fish products or food supplements, and consumption of contaminated drinking water (manganelli et al., 2012). the world health organization (who) set a provisional guideline limit of 1 µg l–1 of mc-lr in drinking water (who, 1998). negative health outcomes resulting from drinking of water contaminated with cyanobacteria or cyanotoxins have been reported worldwide (bláha et al., 2009; van apeldoorn et al., 2007; wood, 2016). the only documented case of cyanobacteria-associated poisoning in africa has been reported from harare, zimbabwe, where annual outbreaks of gastroenteritis among infants occurred after development of cyanobacteria blooms of microcystis aeruginosa (kützing) kützing and dolichospermum flos-aquae (bresson ex bornet & flauhault) in lake chievero (zilberg, 1966). however, there is also a documented case of threetime rise of gastroenteritis and 4.3-time increase of liver cancer incidence rate in harare during the period 19902001 (ndebele and magadza, 2006). although the extent to which this situation is linked to algal toxins is unclear, johansson and olsson (1998) reported that mc concentration in lake chievero was around 13.9 µg l–1 and mc was also detected in municipal tap water. in the last decades, mc occurrence has been investigated in some african countries, with mcs reported from northern africa (algeria, egypt, morocco, tunisia), eastern africa (ethiopia, kenya, mozambique, tanzania, uganda), southern africa (botswana, lesotho, south africa) (mowe et al., 2014; harke et al., 2016; ndlela et al., 2016). however, available data about cyanotoxin contamination are still very scarce for most african countries and being nearly absent for regions such as west and central africa (mowe et al., 2014; harke et al., 2016; ndlela et al., 2016). these are two large geographical areas with 26 countries and nearly half a billion inhabitants, where central africa is represented by nine countries populated by approximately 155 mil people, and west tropical africa by 17 countries with a combined population of about 344 mil people (unsd, 2014). toxic or potentially toxic cyanobacterial blooms seem to occur frequently in this geographical area (addico et al., 2006, 2009, 2017; akin-oriola et al., 2006; berger et al., 2006; haande et al., 2007; mhlanga et al., 2006; odokuma and isirima, 2007). nevertheless, mcs in this region have been so far reported from nigeria (chia et al., 2009a, 2009b; chia and kwaghe, 2015) and detected in water reservoirs brimsu, kwanyarko, kpong and weija in ghana (addico et al., 2006; addico et al., 2017). studies investigating cyanotoxins in drinking water and their removal during the water treatment have been even more scarce on the entire african continent, known to be conducted for example in egypt (mohamed and carmichael, 2000), algeria (nasri et al., 2004), south africa (harding et al., 2009), and recently in two drinking reservoirs in central ghana (addico et al., 2017). in the present study, we investigated seasonal occurrence and removal of cyanobacteria and mcs in four treatment plants supplying the two major metropolitan areas in ghana with drinking water. the study provides very rare but important information regarding the efficiency of drinking water treatment, concentrations and health risks of mcs in drinking water in the understudied geographical region of central and west tropical africa. methods study area the barekese reservoir is a mesotrophic reservoir (addico et al., 2009), which lies on latitude 6° 49′ 50.2″ n and longitude 1° 43′ 21.8″ w (fig. 1). this reservoir was formed in 1970, it has a surface area 6.4 km2 and maximal depth 15 m (amuzu, 1975). it is located on the river ofin, which flows through many farming areas before reaching the dam site (kumasi et al., 2011). the owabi reservoir, also mesotrophic (addico et al., 2009), is located on latitude 6° 44′ 35.7″ n and longitude 1° 42′ 13.4″ w (fig. 1). it was constructed in 1928 and upgraded in 1954. the reservoir has a surface area about 3.5 km2 and a mean depth 7 m (akoto et al., 2014). the owabi reservoir is fed by seven rivers/streams, all of which flow through the densely populated kumasi metropolitan area and the central business and industrial areas. the owabi and barekese reservoirs are both situated in the ashanti region of ghana and serving as major water supplies to kumasi metropolitan area with a population over 2 mil people. the owabi reservoir is designed to produce up to 20% of the total potable water requirement (akoto et al., 2014), while the barekese treatment plant is providing about 80% of the total piped drinking water to the kumasi metropolis (kumasi et al., 2011). the kpong reservoir, described as mesotrophic (addico et al., 2009), is located in the eastern region of ghana on 6° 07′ 1.3″ n 0° 07′ 31.6″ e (fig. 1). it was constructed in 1981 on the volta river system mainly to provide hydroelectricity to supplement power generated from the volta river dam. it has a total surface area of 38 km2, maximal depth of 15 m with a mean depth of 5 m, and a mean annual flow 1183 m3 s–1 (ansa-asare and ansongno nco mm er cia l u se on ly g.n.d. addico et al.94 asante, 1998; quarcoopome et al., 2011). the kpong reservoir apart from power generation is also used for drinking water production, irrigation, recreation and also well known for its fisheries, especially the tilapias. the weija reservoir, a eutrophic reservoir (addico et al., 2009), is situated in the greater accra region of ghana and lies on latitude 5° 34′ 7.1″ n and 0° 20′ 44.8″ w (fig. 1). it has a surface area of about 38 km2, a mean depth of 5 m and a mean annual flow of 54.2 m3 s–1 (ansaasare and ansong-asante, 1998, asante et al., 2008). the weija reservoir was built in 1977 on the densu river system. this river system is under intensive threat from heavy pollution mainly from domestic and agricultural wastes. major crops include maize, cassava, pineapples, pawpaw, banana, sugar cane and vegetables. fishing is also very intensive in the reservoir sometimes with the use of chemicals. the kpong and weija reservoirs are the two main drinking water supplies serving the accratema metropolitan area with a population about 2.3 mil people, with the kpong treatment plant providing approximately 47% and the weija plant about 53% of piped drinking water for the metropolis (stoler et al., 2012). drinking water treatment procedure the water treatment procedure in the studied treatment plants starts from the water intake. in most reservoirs, raw water was collected from depths between 5 to 7 m (addico et al., 2006), with the exception of the barekese plant, where the intake point is placed at the level 1.5-3 m from the surface (amuzu, 1975). raw water is sieved using a mesh to remove big objects like plant parts, twigs etc. this step is followed by flocculation using aluminium sulphate (alum) at a 100 mg l–1 dose, mixing and passing fig. 1. map of the reservoirs and treatment plants under the study. the barekese and owabi reservoirs are located in the ashanti region of ghana and supplying kumasi metropolitan area with drinking water. the kpong reservoir is located in eastern region of ghana, and the weija reservoir in greater accra region, both reservoirs are providing drinking water for accra-tema metropolitan area. no nco mm er cia l u se on ly microcystins in drinking water in ghana 95 through baffles to maximise contact time. the flocs are allowed to settle out of the water in sedimentation tanks. the exception is represented by the kpong reservoir water treatment plant, where alum flocculation is not regularly applied during water treatment and water is prechlorinated before the filtration step. filtration is then done by the rapid sand filtration method and the ph is adjusted to between 6.6 and 8.5 using lime. the final stage involves chlorination employing chlorine gas or calcium hypochlorite with a concentration of 0.5 to 1 mg l–1 of residual chlorine after a contact time of about 30 min (fig. 2). all samples from the water intake to chlorination step were collected consecutively and within the same day. it is important to mention that during one time of sampling at the owabi treatment plant, algaecide treatment with copper sulphate was simultaneously being applied in the reservoir. sampling and cyanobacteria determination samples for mc and cyanobacteria analysis were collected monthly from january-june 2005 from drinking water treatment plants at the barekese and owabi reservoirs, and biweekly at the weija and kpong reservoirs. water samples (1 l) were collected into clean plastic (pet) bottles from each treatment stage, namely: i) raw water at the intake; ii) flocculation (except the kpong treatment plant); iii) sedimentation tanks or clarifiers; iv) filtered water; and v) final chlorinated water. a total number of 127 samples were analysed for intracellular mcs in the four reservoirs, whilst 59 samples were analysed for dissolved mcs. the lower number of samples analysed for dissolved mcs was due to financial constraints during the field work in ghana and losses during the sample transport from ghana to the united kingdom. samples for microscopic determination of cyanobacterial species composition were collected from raw water and final treated water using plankton net (25 µm) or by simply filling a bucket. net samples were preserved for taxonomic work with formalin at the final concentration of 2% (v/v), whilst water samples were preserved in lugol’s solution for quantitative microscopical analysis as described in addico et al. (2006, 2009) using olympus bx51 and bx60 microscopes equipped with objectives 10, 20, 40, 60 and 100x (olympus). briefly, the aliquots of the samples were transferred into counting chambers for analysis, where all colonies and filaments were counted as individuals. the average number of cells was determined for 20 individuals and cell concentration was calculated. sub-samples for counting picocyanobacteria were filtered through a 0.2 μm nucleopore filter prestained with irgalan black. cells were stained with dapi (4-diamidino-2-phenylindole dihydrochloride) and counted under the fluorescence (excitation 330-385 nm, emission 510-560 nm). about 300-400 picocyanobacterial cells were counted for each sample. all data on abundance were expressed as number of cells per ml, including the cells inside colonies. identification of cyanobacteria species was carried out at the institute of botany of the czech academy of sciences, trebon, czech republic under the supervision prof. jiří komárek and using the recent taxonomical literature (komárek and anagnostidis, 1999, 2005). extraction of cell-bound (intracellular) mcs water samples (1 l) were filtered through preweighed gf/c filter (1.2 µm mesh, whatman). the cells collected on the filters were frozen overnight and freeze-dried. freeze-dried cells on filters were stored at -20°c until extracted for hplc analysis. extraction of cell-bound (intracellular) toxins from freeze-dried cells was done as described by harada et al. (1999). cells were extracted with 20 ml of 75% aqueous methanol (fastner et al., 1998) for 1 hour. this extraction step was repeated three times, the extracts from the individual steps were combined and then dried using a rotary evaporator. the concentrated extract was dissolved in 400 µl methanol prior to hplc analysis, filtered through 0.45 µm nylon syringe filter (millipore). fig. 2. summary of drinking water treatment process employed in ghanaian plants barekese, owabi, kpong and weija. no nco mm er cia l u se on ly g.n.d. addico et al.96 extraction of dissolved (extracellular) mcs filtrates of water samples (1 l) filtered through gf/c filter (see above) were processed according to harada et al. (1999). briefly, filtrates were treated with sodium thiosulphate (2 mg l–1), acidified with trifluoroacetic acid (tfa, 0.1%, v/v) and concentrated using solid phase extraction by ods cartridges (supelclean lc-18, 3 ml tube, supelco). cartridges were activated with 5 ml of methanol and rinsed with 5 ml of distilled water prior to the application of the sample. mcs were then eluted with 15 ml of 0.1% tfa in methanol, the eluate was evaporated to dryness by rotary vacuum evaporation (45 c) and then redissolved in 400 µl methanol in an ultrasonic bath. identification and quantification of mcs mcs were identified and quantified using high performance liquid chromatography (hplc agilent 1100 series) system, coupled with a diode array detector (dad). mcs were separated on a c-18 column luna 150×4.60 mm, 5 µm (phenomenex) at 30°c using a flow rate of 1 ml min–1. the binary gradient of the mobile phase consisted of (a) h2o+0.05% tfa and (b) acetonitrile +0.05% tfa, with a linear increase from 30 to 70% b between 0-30 min. the injection volume was 20 µl. chromatograms were recorded at 238 nm. uv spectra (200 to 300 nm) of all chromatographic peaks were carefully checked and compared to the spectra of mc standards: mc-lf, -lr, -lw, -rr and -ly (alexis biochemicals). peaks possessing the uv spectrum characteristic for mcs were quantified using a calibration curve (n=5, r2=0.999) of the corresponding standard with the matching retention time. unidentified peaks possessing the uv spectrum characteristic for mcs but not matching the retention time of the standards were quantified as mc-lr equivalents using the calibration curve of mc-lr (mcelhiney and lawton, 2005). the detection limit of the method (lod) was 0.01 µg l–1 for the individual mc variant. results cyanobacteria removal in this study, we complemented previous data on cyanobacterial concentrations in the raw water (addico et al., 2009) with a new data set on cyanobacterial cell counts in the final treated water, and also with mc analyses, in order to discuss relationships between mc occurrence, cyanobacterial diversity, and their removal during the drinking water treatment. as reported, all four reservoirs were dominated by cyanobacteria, which accounted for 70-90% of phytoplankton biomass (addico et al., 2009). detailed results of microscopical analyses of cyanobacterial species composition are summarized in tabs. 1-4, the complete list of the identified species is provided in the supplementary tab. 1. representatives of picocyanobacterial genera cyanogranis, aphanocapsa and geitlerinema were among the most aboundant species tab. 1. concentration of cyanobacterial cells (cell ml–1) in the water intake and in the final treated water at the barekese drinking water treatment plant during the jan-may 2005. barekese january february march april may average intake* final intake* final intake* final intake* final intake* final intake final removal (%) anabaena austro-africana 0 0 35 0 140 0 89 0 926 0 238 0 100.0 anabaena nygaardii 7768 0 9010 54 11,509 10 4923 0 1922 0 7026 13 99.8 chroococcus cronbergae 756 0 467 0 899 0 1281 0 874 0 855 0 100.0 cyanogranis ferruginea 201,870 98 229,018 1143 191,002 475 48,594 45 125,000 87 159,097 370 99.8 cylindrospermopsis raciborskii 1007 11 4005 28 2086 12 4272 0 2760 0 2826 10 99.6 merismopedia punctata 2987 0 1998 0 1254 0 995 0 1075 0 1662 0 100.0 merismopedia tenuissima 7098 0 5998 0 5990 0 3709 0 2136 0 4986 0 100.0 microcystis aeruginosa 501 0 429 0 557 0 400 0 566 0 491 0 100.0 oscillatoria princeps 5783 0 6602 0 4998 0 5340 0 7251 0 5995 0 100.0 planktolyngbya minor 3056 0 2955 0 1565 0 1427 0 5073 0 2815 0 100.0 planktothrix lacustris var. solitaria 2008 3 3090 14 3163 8 5146 14 2895 25 3260 13 99.6 planktothrix sp. 98 33 801 99 2675 10 1226 26 3925 34 1745 40 97.7 pseudanabaena recta 6780 0 6675 10 3727 4 3888 7 925 2 4399 5 99.9 radiocystis fernandoi 0 0 96 0 230 0 0 0 431 0 151 0 100.0 romeria elegans 56 0 0 0 18 0 53 1 36 5 33 1 96.3 total (cell ml–1) 239,768 145 271,179 1,348 229,813 519 81,343 93 155,795 153 229,813 153 removal (%) 99.9 99.5 99.8 99.9 99.9 99.9 *data adapted from addico et al. (2009). no nco mm er cia l u se on ly microcystins in drinking water in ghana 97 within the cyanobacterial communities in the studied reservoirs (tabs. 1-4). in the barekese and owabi reservoirs, which are located in the same ecological zone in the ashanti region (fig. 1), cyanogranis ferruginea (f. wawrik) hindak ex hindak accounted for the majority of cyanobacterial cells. c. ferruginea population in the raw water at the barekese treatment plant ranged between 60-85% of total cyanobacteria cell counts (tab. 1). this species was the most abundant in the treated water as well, accompanied also with planktothrix agardhii (gomont) k. anagnostidis & j. komárek, planktothrix lacustris (klebahn) i. umezaki & m. watanabe, and eventually by cylindrospermopsis raciborskii (woloszynska) seenayya & subba raju, pseudanabaena recta komárek & cronberg, anabaena nygaardii cronberg & komárek (tab. 1). concentrations of cyanobacterial cell in the final water from the barekese reservoir ranged between 93-1,348 cell ml–1. in the owabi treatment plant, c. ferruginea represented 95-97% of the total cyanobacteria cell counts (tab. 2). in addition to c. ferruginea, cyanobacteria aphanocapsa holstatica (lemmermann) g. cronberg & komárek, p. recta, and leptolyngbya sp. were detected most frequently in the treated water from owabi, with total cyanobacterial counts between 95-1099 cells ml–1 (tab. 2). geitlerinema unigranulatum (c. agardh ex gomont) anagnostidis was the most abundant cyanobacterium in the kpong reservoir, representing 65-78% of total cyanobacterial cell counts in the raw water samples (tab. 3). however, p. agardhii was in average the most abundant species found in the final water from the kpong reservoir, followed by g. unigranulatum and c. raciborskii, while other species were detected in the treated water only occassionally. total cyanobacterial cell counts in the final water were between 173-845 cell ml–1 during the sampling period (tab. 3). cyanobacterial community in the weija reservoir was the most diverse one (tab. 4), when the most abundant cyanobacterial species aphanocapsa nubilum komárek & h.j. kling accounted only for 18-26% of total cyanobacterial cell counts throughout six months of sampling, while being accompanied with merismopedia tenuissima lemmermann (14-19%), planktolyngbya minor (geitler & ruttner) komárek & cronberg (815%), p. recta (6-18%) and others. the most abundant species in the treated water was chroococcus cronbergae j. komárek & e. novelo, which penetrated into the final stage throughout the study, along with a. nubilum, m. aeruginosa, a. nygardii, p. agardhii and c. raciborskii. the cyanobacterial cell counts in the final water from the weija reservoir were found to be between 369-3,055 cell ml–1 (tab. 4). overall, the drinking water treatment process eliminated >97-99.9% of cyanobacterial cells, however, cyanobacteria were detected in 100% samples of treated water collected from all four treatment plants during the entire sampling period. mc removal commonly occurring mc variants identified in the examined reservoirs were mc-lr, -lf, -rr and -yr. the highest diversity of mc variants was observed in the weija reservoir, where also two additional peaks possessing mclike uv absorption spectrum were identified. out of the 26 samples of raw water, 17 samples (65%) contained intracellular mcs (fig. 3, tab. 5). during the water treatment process, concentrations of both intracellular as well as extracellular toxins generally decreased with the treatment tab. 2. concentration of cyanobacterial cells (cell ml–1) in the water intake and in the final treated water at the owabi drinking water treatment plant during the jan-may 2005. owabi january february march april may average intake* final intake* final intake* final intake* final intake* final intake final removal (%) anabaena nygaardii 65 0 0 0 0 0 16 0 0 0 16 0 100.0 aphanocapsa holsatica 2092 76 1980 66 2541 97 1997 45 2672 76 2256 72 96.8 chroococcus cronbergae 67 0 0 0 10 25 0 0 0 0 15 5 67.5 cyanogranis ferruginea 225,317 34 220,001 47 165,002 901 157,005 37 278,430 98 209,151 223 99.9 cylindrospermopsis raciborskii 24 0 15 0 24 0 45 0 23 0 26 0 100.0 leptolyngbya sp. 946 0 882 27 98 0 107 13 905 9 588 10 98.3 merismopedia tenuissima 62 0 77 0 102 0 91 0 75 0 81 0 100.0 planktolyngbya limnetica 772 0 1372 0 1532 0 2109 0 815 0 1320 0 100.0 planktolyngbya minor 2008 0 1247 0 2349 0 2129 0 3761 0 2299 0 100.0 pseudanabaena recta 1465 56 2165 35 1645 76 1705 0 868 20 1570 37 97.6 total (cell ml–1) 232,818 166 227,739 175 173,303 1099 165,204 95 287,549 203 227,739 175 removal (%) 99.9 99.9 99.4 99.9 99.9 99.9 *data adapted from addico et al. (2009). no nco mm er cia l u se on ly g.n.d. addico et al.98 step in all four individual treatment plants (fig. 3). statistically significant (p<0.05) correlation (spearman’s rank correlation coefficient ρ) between mc concentration and the order of treatment step was found: ρ=0.996 (barekese), ρ=0.861 (owabi), ρ=0.987 (kpong) and ρ=0.899 (weija) for intracellular toxins, and ρ=1 (barekese), ρ=0.911 (owabi), ρ=0.965 (kpong) and ρ=0.980 (weija) for dissolved toxins. however, increases in concentration of both intracellular and dissolved mc were observed in some cases after the flocculation/sedimentation steps of water treatment (fig. 3 and supplementary figs. 1-4). five samples of the treated water (17%) contained intracellular or fig. 3. combined data on mc concentrations at different stages of four drinking water treatment plants in ghana during jan-jun 2005. boxes plot median values (middle lines), 25th and 75th percentils (boxes), 10th and 90th percentils (error bars) and outliers (circles). ratio between median intracellular (ic) and dissolved (dis) mc concentrations was calculated for different treatment steps and ploted as a line graph. hash indicates significant difference between concentration of ic and dis mc at the particular step of drinking water treatment (p<0.05, mann-whitney test). asterisks indicate significant difference between mc concentration in the water intake and a particular treatment step (p<0.05, mann-whitney test). values below the method lod (0.01 µg l–1) were susbstituted with lod/2. tab. 3. concentration of cyanobacterial cells (cell ml–1) in the water intake and in the final treated water at the kpong drinking water treatment plant during the jan-may 2005. kpong january february march april may average intake* final intake* final intake* final intake* final intake* final intake final removal (%) chroococcus cronbergae 690 0 382 0 656 0 609 0 541 0 575 0 100.0 coelomoron tropicale 0 0 13 0 0 0 0 0 40 0 11 0 100.0 cylindrospermopsis cuspis 1226 0 1232 0 2401 0 1003 11 3980 0 1968 2 99.9 cylindrospermopsis raciborskii 2625 30 1806 39 2216 87 3873 35 2082 60 2520 50 98.0 geitlerinema unigranulatum 31,584 127 39,562 89 49,325 179 30,252 389 33,546 16 36854 160 99.6 merismopedia punctata 1204 0 924 0 840 0 550 0 1082 0 920 0 100.0 merismopedia tenuissima 2033 0 3693 0 2991 0 1563 0 3865 0 2829 0 100.0 planktolyngbya minor 55 0 61 0 90 0 61 0 1531 67 359 13 96.3 planktothrix agardhii 2475 65 4123 45 4070 293 2846 410 4352 41 3573 171 95.2 pseudanabaena recta 673 0 904 0 719 0 646 0 320 0 652 0 100.0 total (cell ml–1) 42,562 222 52,698 173 63,306 559 41,400 845 51,336 184 51,336 222 removal (%) 99.5 99.7 99.1 98.0 99.6 99.6 *data adapted from addico et al. (2009). no nco mm er cia l u se on ly microcystins in drinking water in ghana 99 particle-associated mcs (maximal detected concentration was 0.61 µg l–1), and two samples of the treated water (14%) contained dissolved mcs at concentrations 0.57 µg l–1 (kpong) and 0.81 µg l-1 (weija) (tab. 5). concentrations of intracellular toxins significantly correlated with the concentrations of the cyanobacterial cells in the treated water (ρ=0.561, p<0.01). in the barekese reservoir, mcs were detected during two out of six sampling months (tab. 5, supplementary fig. 1). in one instance (february 10th), intracellular mcs were found in the sedimentation step and then in the sample of treated water (0.45 µg l–1), while dissolved mcs were found in the flocculation stage (supplementary fig. 1). in april, intracellular mcs were detected in one sample of raw water at the concentration 0.46 µg l–1, which further increased in the flocculation stage, but then decreased below the detectable levels in the next treatment step (tab. 5, supplementary fig. 1). the owabi reservoir was found to be more contaminated with mcs. all raw water samples from the owabi treatment plant contained intracellular mcs (tab. 5, supplementary fig. 2). the highest detected concentration of intracellular mcs in the intake water from the owabi reservoir was 8.73 µg l–1, and intracellular toxins were detected also in the final water in one instance (0.07 µg l–1, march 17th). dissolved mcs could be found in samples from all treatment stages of the owabi treatment plant with the exception of the final stage (tab. 5, supplementary fig. 2). in the kpong reservoir, mcs were detected relatively less frequently. intracellular toxins were found only in two out of seven samples of intake water (tab. 5, supplementary fig. 3). however, the kpong drinking water treatment plant, which does not have a flocculation stage, had two out of eight samples contaminated with intracellular mcs at the final chlorination stage (0.13 and 0.46 µg l–1, march tab. 4. concentration of cyanobacterial cells (cell ml–1) in the water intake and in the final treated water at the weija drinking water treatment plant during the jan-may 2005. weijajanuary february march april may average intake* final intake* final intake* final intake* final intake* final intake final removal (%) anabaena austro-africana 398 0 330 0 1,075 0 961 0 2807 0 1,114 0 100.0 anabaena nygaardii 2104 21 7080 475 5173 67 6485 35 11,297 15 6428 123 98.1 anabaenopsis ambigua 177 0 763 0 738 0 792 0 45 0 503 0 100.0 anabaenopsis tanganyikae 34 0 109 69 60 0 879 64 557 0 328 27 91.9 aphanocapsa holsatica 6302 25 2677 0 821 0 2151 0 6350 0 3660 5 99.9 aphanocapsa nubilum 24,567 12 20,538 883 20,832 80 26,883 42 19,511 67 22,466 217 99.0 chroococcus cronbergae 3141 515 5398 515 5284 962 5989 483 4586 224 4880 540 88.9 coelomoron tropicale 52 0 20 0 0 0 0 0 0 0 14 0 100.0 cyanogranis ferruginea 0 0 80 0 1208 0 903 0 1773 0 793 0 100.0 cylindrospermopsis cuspis 10 23 12 14 96 0 76 0 87 0 56 7 86.8 cylindrospermopsis raciborskii 1284 17 5148 97 4565 41 5051 25 4112 12 4032 38 99.0 geitlerinema unigranulatum 8 0 6 0 0 0 0 0 9 0 5 0 100.0 lyngbya sp. 6 0 0 0 5 0 6 0 12 0 6 0 100.0 merismopedia punctata 1414 0 627 0 263 0 3006 0 1461 0 1354 0 100.0 merismopedia tenuissima 20,859 48 20,907 34 17,794 0 15,059 5 12,916 0 17,507 17 99.9 microcystis aeruginosa 692 53 1782 744 2958 66 3051 47 2183 33 2133 189 91.2 microcystis viridis 0 0 0 0 0 0 0 0 493 0 99 0 100.0 microcystis wesenbergii 250 0 0 0 562 0 0 0 318 0 226 0 100.0 planktolyngbya circumcreta 162 0 75 0 416 47 439 0 219 0 262 9 96.4 planktolyngbya limnetica 3645 0 2774 0 5060 0 1843 0 2220 0 3108 0 100.0 planktolyngbya minor 13,843 0 16,883 0 12,610 0 11,736 0 7069 0 12,428 0 100.0 planktothrix agardhii 2395 11 2903 224 1311 37 2181 66 1757 18 2109 71 96.6 planktothrix lacustris var. solitaria 8723 0 6873 0 5299 0 2443 0 6130 0 5893 0 100.0 pseudanabaena recta 13,446 0 15,838 0 18,835 0 9624 0 5058 0 12,560 0 100.0 radiocystis fernandoi 3992 0 3395 0 2516 0 3089 0 1330 0 2864 0 100.0 romeria elegans 0 0 20 0 8 0 1 0 10 0 8 0 100.0 total (cell ml) 107,500 725 114,231 3055 107,485 1300 102,643 767 92,304 369 107,485 767 removal (%) 99.3 97.3 98.8 99.3 99.6 99.3 *data adapted from addico et al. (2009). no nco mm er cia l u se on ly g.n.d. addico et al.100 ta b. 5 . s um m ar y of m c a na ly se s fr om d if fe re nt tr ea tm en t s te ps o f t he fo ur g ha na ia n tr ea tm en t p la nt s. in tr ac el lu la r m c s (µ g l –1 ) d is so lv ed m c s (µ g l –1 ) (r es er vo ir t re at m en t n um be r of s am pl es c on ce nt ra ti on m ed ia n n um be r of s am pl es c on ce nt ra ti on m ed ia n (s am pl in g pe ri od ) s te p to ta l >l o d > 1 µg l –1 ra ng e t ot al > l o d > 1 µg l –1 r an ge b ar ek es e in ta ke 4 1 (2 5% ) 0 < l o d -0 .4 6 < l o d 3 0 0 < l o d < l o d (1 0 ja n7 ju n) f lo cc ul at io n 5 1 (2 0% ) 1 (2 0% ) < l o d -1 5. 50 <l o d 3 1 (3 3% ) 0 < l o d -1 .0 0 <l o d s ed im en ta tio n 6 1 (1 7% ) 1 (1 7% ) < l o d -1 .6 5 < l o d 3 0 0 < l o d < l o d f ilt er ed 5 0 0 <l o d < l o d 3 0 0 < l o d < l o d c hl or in at io n 6 1 (1 7% ) 0 < l o d -0 .4 5 < l o d 2 0 0 < l o d < l o d o w ab i in ta ke 4 4 (1 00 % ) 2 (5 0% ) 0 .0 88. 73 0 .7 0 1 1 (1 00 % ) 0 0. 60 0 .6 0 (1 0 ja n7 ju n) f lo cc ul at io n 5 3 (6 0% ) 1 (2 0% ) < l o d -6 .2 4 0. 09 2 2 (1 00 % ) 0 0 .2 90. 87 0 .5 8 s ed im en ta tio n 4 3 (7 5% ) 1 (2 5% ) < l o d -4 .0 2 0. 12 2 2 (1 00 % ) 0 0 .0 90. 32 0 .2 1 f ilt er ed 6 2 (3 3% ) 0 < l o d -0 .4 5 < l o d 4 2 (5 0% ) 0 < l o d -0 .8 9 0 .0 8 c hl or in at io n 5 1 (2 0% ) 0 < l o d -0 .0 7 < l o d 4 0 0 < l o d < l o d k po ng i nt ak e 7 2 (2 9% ) 1 (1 4% ) < l o d -4 .6 9 < l o d 3 1 (3 3% ) 0 < l o d -0 .1 0 <l o d (1 3 ja n10 j un ) s ed im en ta tio n 6 1 (1 7% ) 0 < l o d -0 .8 8 < l o d 5 3 (6 0% ) 0 < l o d -0 .9 2 0 .1 5 f ilt er ed 6 2 (3 3% ) 0 < l o d -0 .1 2 < l o d 4 1 (2 5% ) 0 < l o d -0 .2 0 <l o d c hl or in at io n 8 2 (2 5% ) 0 < l o d -0 .4 6 < l o d 4 1 (2 5% ) 0 < l o d -0 .5 7 <l o d w ei ja in ta ke 1 1 1 0 (9 1% ) 4 (3 6% ) < l o d -8 .5 6 0. 76 2 1 (5 0% ) 0 < l o d -0 .0 3 0 .0 2 (1 j an -1 4 ju n) f lo cc ul at io n 1 1 9 (8 2% ) 5 (4 5% ) < l o d -1 0. 92 0 .6 8 5 2 (4 0% ) 0 < l o d -0 .6 7 <l o d s ed im en ta tio n 9 4 (4 4% ) 1 (1 1% ) < l o d -3 .5 3 < l o d 3 2 (6 7% ) 1 (3 3% ) < l o d -1 .8 2 0 .0 1 f ilt er ed 9 1 (1 1% ) 0 < l o d -0 .3 6 < l o d 2 0 0 < l o d < l o d c hl or in at io n 10 1 (1 0% ) 0 < l o d -0 .6 1 < l o d 4 1 (2 5% ) 0 < l o d -0 .8 1 <l o d to ta l in ta ke 2 6 1 7 (6 5% ) 7 (2 7% ) < l o d -8 .7 3 0. 36 9 3 (3 3% ) 0 < l o d -0 .6 0 <l o d f lo cc ul at io n 2 1 1 3 (6 2% ) 7 (3 3% ) < l o d -1 5. 50 0 .3 2 1 0 5 (5 0% ) 0 < l o d -1 .0 0 0 .1 5 s ed im en ta tio n 25 9 (3 6% ) 3 (1 2% ) < l o d -4 .0 2 < l o d 13 7 (5 4% ) 1 (8 % ) <l o d -1 .8 2 0 .0 5 f ilt er ed 26 5 (1 9% ) 0 < l o d -0 .4 5 < l o d 13 3 (2 3% ) 0 < l o d -0 .8 9 <l o d c hl or in at io n 29 5 (1 7% ) 0 < l o d -0 .6 1 < l o d 14 2 (1 4% ) 0 < l o d -0 .8 1 <l o d t ot al 12 7 4 9 (3 9% ) 1 7 (1 3% ) <l o d -1 5. 50 <l o d 59 2 0 (3 4% ) 1 (2 % ) <l o d -1 .8 2 <l o d l o d : 0 .0 1 µg l –1 , t he v al ue s <l o d w er e su bs tit ut ed w ith l o d /2 fo r m ed ia n ca lc ul at io n. no nco mm er cia l u se on ly microcystins in drinking water in ghana 101 10th and april 8th). concentrations of dissolved mcs showed a remarkable trend on march 24th, when increased from 0.1 µg l–1 at the water intake to 0.57 µg l–1 at the final drinking water treatment stage (supplementary fig. 3). the raw water from the weija reservoir was more heavily contaminated with mcs, which were found in the particle fraction in ten out of 11 samples at the water intake, with the highest detected concentration 8.56 µg l–1 (tab. 5, supplementary fig. 4). concentrations of intracellular mcs in the weija reservoir often increased at the flocculation step and gradually decreased as the treatment proceeded to the final water stage (supplementary fig. 4). one sample of the final drinking water contained 0.61 µg l–1 of intracellular mcs (february 9th). the highest concentration of dissolved mcs in the treated water (0.81 µg l–1) was observed on january 24th, and resulted from an increase in the concentrations of dissolved toxins during the treatment process (supplementary fig. 4). discussion as recently reviewed by bernard et al. (2017), ten cyanobacterial genera detected in our study (tabs. 1-4, supplementary tab. 1) have been previously associated with mc production: anabaena/dolichospermum, anabaenopsis, aphanocapsa, geitlerinema, leptolyngbya, merismopedia, microcystis, oscillatoria/planktothrix, pseudanabaena and radiocystis. indeed, we frequently detected intracellular mcs in the samples of raw water collected from four ghanaian drinking water treatment plants. the cyanobacteria composition and potential mcproducers differed among the sampled reservoirs. the most abundant species in the owabi reservoir was c. ferruginea, which has not been previously reported to produce mcs and little information on its ecology or toxicity exists (addico et al., 2009). since mcs were detected with high frequency in samples from the owabi reservoir, it might suggest this cyanobacterium could produce mcs. however, representatives of the above mentioned cyanobacterial genera associated with mc production, namely a. nygaardii, a. holsatica, leptolyngbya sp. and p. recta were also present in the owabi samples. although concentrations of these species were much lower than c. ferruginea cell counts, their share of the total biomass might have been relatively larger due to the differences in their cell size. since a. holstatica and p. recta (and also c. cronbergae) were detected along with c. ferruginea in the only owabi sample of the treated water positive for mcs, it does not rule out these species as putative mc producers. in addition, mcs were detected much less frequently in the barekese reservoir, where c. ferruginea occurred in concentrations similar to owabi. other representatives of potentially mc-producing genera, such as a. nygaardii, planktothrix sp., p. lacustris var. solitaria and p. recta were also present in the barekese reservoir, including the sample of treated water positive for intracellular mcs (february 10th). thus, other species might have been responsible for mc production in the two reservoirs with abundant population of c. ferruginea, and experiments with the isolated strains are required to confirm picocyanobacterium c. ferruginea as a mc producer. the highest encounted species in the kpong reservoir was g. unigranulatum. cyanobacteria from geitlerinema genus have been recently reported to produce mcs (bernard et al., 2017; myers et al., 2007; valdor and aboal 2007). however, other putative producers of mcs were also present in the kpong reservoir, most notably p. agardhii, which also penetrated into the final stage of the water treatment on the sampling days when intracellular mcs were detected in the final water (march 10th, april 8th). the weija reservoir had a high diversity of cyanobacteria, including anabaenopsis tanganyikae (g. s. west) woloszynska & v. v. miller, a. nygaardii, a. nubilum, m. tenuissima, m. aeruginosa, p. agardhii or p. recta, which are not only species representing the genera potentially producing mcs (bernard et al., 2017), but they were also detected in the final treated water containing intracellular mcs (february 9th). interestingly, the major cyanobacterial species in the weija and kpong reservoirs changed from years 20022003, when m. aeruginosa and d. flos-aquae were the most abundant, with their average concentrations in the water intake were around 10,000 cell ml–1 and corresponded to 41-51% of the total cyanobacterial cell counts (addico et al., 2006). addico et al. (2006) also observed breakthrough of cyanobacteria to the final treated water at the averaged total cell counts 685 cell ml–1, which is comparable to our current study. at that time, mcs were analyzed only in the water intake, where present at concentrations 0.03 µg l–1 in the kpong reservoir and 3.21 µg l–1 in the weija reservoir (addico et al., 2006). in 2010, cyanobacteria and mcs were analyzed also in the brimsu and kwanyarko treatment plants in ghana. p. agardhii, d. flos-aque and p. recta were the most abundant species in the water intake from the brimsu reservoir, whereas m. aeruginosa followed by p. agardhii and merismopedia punctata meyen accounted nearly for 80% of the cyanobacterial cell counts in the raw water in the kwanyarko treatment plant (addico et al., 2017). concentrations of cyanobacteria in the water intake were much lower (brimsu: 2183 cell ml–1, kwanyarko: 1286 cell ml–1) than in the current study or in the study of addico et al. (2006), while concentrations of mcs in raw water were 0.79 µg l–1 in brimsu and 0.10 µg l–1 in kwanyarko. however, cyanobacteria (brimsu: 103 cell ml–1, kwanyarko: 85 cell ml–1) as well as mcs (0.10 µg l–1 in both plants) were detected also in the final water (addico et al., 2017). no nco mm er cia l u se on ly g.n.d. addico et al.102 these previously published findings together with the results presented in this paper demonstrate that mc-producing species and mc contamination are common and enduring health hazard in drinking water reservoirs in ghana. concentrations of cyanobacteria and mc have been shown to significantly decrease during the treatment process. apparently, conventional drinking water treatment plants existing in ghana were sufficiently effective in removing of mcs to the levels below the who limit for drinking water. however, samples of the final treated water often contained cyanobacterial cells. in addition to the most abundant species, i.e., c. ferruginea, g. unigranulatum and a. nubilum, other cyanobacterial species were observed to infiltrate the final water. these included most notably planktothrix sp., c. raciborskii (barekese, kpong and weija), p. recta (barekese, owabi), a. holstatica, leptolyngbya sp. (owabi), a. nygaardi (barekese, weija), c. cronbergae and m. aeruginosa (weija), which were often found in the final water in disproportionately higher share than in the raw water. correspondingly, intracellular or particle-associated mcs were found in the treated water, which indicates that the treatment process should be further optimized to remove cyanobacterial cells more effectively. this could include adjustments of the flocculant amount or application of a flocculant aid in response to changes in the raw water quality affecting coagulationflocculation-sedimentation process, such as turbidity, concentration of suspended solids and natural organic matter, phytoplankton abundance and composition (hoeger et al., 2005; merel et al., 2013; westrick et al., 2010). our results also indicate that the stage with alum flocculation during the drinking water treatment yielded the higest intracellular mc concentrations. samples from the flocculation stage were taken in the same manner as all the stations (water mixed with flocs). this might be due to large number of cells trapped in the flocs, as reported by pietsch et al. (2002). interestingly, the overall frequency of samples containing dissolved mcs also increased after the flocculation step (barekese, owabi, weija) or pre-chlorination (kpong). this is in agreement with previous observations that these treatment steps can lead to the damage of cyanobacterial cells due to sheer stress or chemical oxidation, resulting in the release of cyanotoxins into the surrounding water (hoeger et al., 2005; hrudey et al., 1999; merel et al., 2013; pietsch et al., 2002; westrick et al., 2010). concentrations of dissolved mcs tended to decrease during the further steps of the water treatment, which might be attributed to their biodegradation during rapid sand filtration (ho et al., 2006) or chlorination (merel et al., 2010). however, the applied water treatment procedures in some instances actually resulted in the increase of dissolved mc concentrations even in the final treated water (weija, january 24th, and kpong, march 24th), which indicates that the conventional treatment procedures are only partially effective for removal of dissolved toxins. among the four reservoirs studied, the owabi reservoir was the most contaminated reservoir in terms of both intracellular and dissolved mcs. as mentioned earlier in the methodology at the time of sampling the owabi treatment plant, algaecide treatment with copper sulphate was simultaneously being done during water treatment and this may account for the higher level of samples containing dissolved mcs in this reservoir, putting consumers at higher risk of exposure to mcs. the dangers of using algaecides in drinking water reservoirs was demonstrated in an incident which occurred on tropical palm island, australia, where members of the community became ill with hepato-enteritis following treatment of the water supply reservoir with copper sulphate after complaints of bad taste and odour from the population (byth, 1980; bourke et al., 1983). jones and orr (1994) suggested that if algaecides are used to control toxic cyanobacteria, the reservoir should be isolated for a period to allow the toxins and odour to degrade. although mc concentrations detected in samples of final treated water in our study were lower (0.07-0.81 µg l–1) than recommended who limit for mc-lr, hoeger et al. (2005) cautioned that due to tumor promoting activities of mcs, chronic exposure of populations to concentration above 0.1 µg l–1 should be avoided. in fact, negative health outcomes have been associated with consumption of drinking water containing less than 1 µg l–1 of mcs, such as high incidence of primary liver cancer associated with the use of mc-contaminated water (maximal concentration 0.46 µg l–1) in southern china (ueno et al., 1996; yu, 1989, 1995). however, it must be mentioned that other risk factors such as hepatitis b and exposure to aflatoxins from maize and other cereals were also implicated (ueno et al., 1996). chronic exposures to water contaminated with mcs at concentrations 1.1-3.28 µg l–1 have been linked with the increased incidence of colorectal cancer (zhou et al. 2002) or liver damage (chen et al., 2009; li et al., 2011) in china. in addition, there are reports on statistically significant associations between cyanobacterial blooms and non-alcoholic liver disease in usa (zhang et al., 2015) or primary liver cancer in serbia (svircev et al., 2009, 2013, 2014). the levels of mcs similar to those observed in this study have been linked also to acute effects (wood, 2016). in sweden, mcs at concentrations 0.1-0.5 µg l–1 were detected in the drinking water during an acute non-fatal outbreak of gastroenteritis associated with mass development of mc-producing p. agardhii in the raw water supply (annadotter et al., 2001). these reports indicate that negative health outcomes resulting from both acute and chronic exposure to drinking water contaminated with cyanobacteria might be associated with concentrations of mcs around 1 µg l–1 which might also interact with other factors, such no nco mm er cia l u se on ly microcystins in drinking water in ghana 103 as other bioactive substances from cyanobacteria (lipopolysaccharides and other cyanobacterial metabolites), pathogens or anthropogenic contaminants. more importantly, about 40% of people in ghana overall and up to 80% in the rural areas does not have an access to the treated piped water (kumasi et al., 2011; stoler et al., 2012), and thus eventually relies on surface water and shallow wells without any treatment. this situation might pose much more serious human health risk, since according to our study, mcs occurred very frequently in the raw water, often at concentrations higher than who limit (27% samples of intracellular toxins). massive proliferations of cyanobacteria are likely to continue during the next years due to the ongoing eutrophication and climate change (o’neil et al., 2012). in combination with unsolved and increasing demands for safe drinking water in ghana (stoler et al., 2012), cyanotoxin occurrence in drinking water reservoirs in ghana represents long-standing and probably intensifying issue, which needs to be adequately addressed by scientists, water treatment management and public health authorities. conclusions safe drinking water has been recognised as one of the most critical factors to guarantee long-term population health. our study clearly demonstrated that the four studied treatment plants successfully met the who guideline value of 1 µg l–1. however, not only dissolved but also intracellular toxins and cyanobacterial cells were found in the final treated water, which indicates a borderline efficiency of the water treatment, possibly resulting in higher toxin concentrations if the treatment procedure gets compromised. in addition, populations drinking untreated surface water can be more often exposed to mc concentrations exceeding 1 µg l–1. the lessons to be learnt from this study are: i) the development of cyanobacterial blooms and mc contamination represent a serious health hazard all over the world and ghana is no exception; ii) since there are neither cyanobacteria monitoring programs nor guidelines for mc concentration in drinking water in ghana, it is recommended that all reservoirs used for drinking water production should be monitored regularly to optimize water treatment process and to protect the public health; iii) other cyanotoxins such as cylindrospermopsin or neurotoxins should be also considered in the further research, since their potential producers were detected in our study; iv) surface water resources should be protected through proper watershed management practices to reduce eutrophication and thus cyanobacteria blooms; v) mass education should also be undertaken to inform the public of the dangers of cyanobacteria and toxins and more importantly about their negative health effects on humans. acknowledgments the research was supported by the government of ghana. microcystin analyses were also supported by a research project no. rvo 67985939, and by czech ministry of education, youth and sports (projects no. lo1214 and lm2015051). we would also like to thank professor jiří komárek (institute of botany, czech academy of sciences, trebon, czech republic) for the help with determination of cyanobacteria, professor linda lawton and dr. christine edwards (robert gordon university, aberdeen, uk) for the help with microcystin analysis, dr. felix akpabey and ms. doreen ofosu (csir water research institute, accra, ghana) for their assistance with sample collection and processing, ralf bubliz and kerstin hanisch (university of hull, hull, uk) for the help with hplc analyses. references addico g, hardege j, komárek j, babica p, de-graftjohnson kaa, 2006. cyanobacteria species identified in the weija and kpong reservoirs, ghana, and their implications for drinking water quality with respect to microcystin. afr. j. marine sci. 28:451-456. addico g, hardege j, komárek j, de-graft-johnson kaa, 2009. cyanobacterial diversity and biomass in relation to nutrient regime of four freshwater reservoirs sourced for the production of drinking water in ghana. algol. stud. 130:81-108. addico g, lawton l, edwards c, 2017. hepatotoxic-microcystins in two drinking water reservoirs in the central region of ghana. toxicol. forensic med. 2:1-11. akin-oriola ga, anetekhai ma, oriola a, 2006. algal blooms in nigerian waters: an overview. afr. j. marine sci. 28:219-224. akoto o, gyamfi o, darko, g, barnes vx, 2014. changes in water quality in the owabi water treatment plant in ghana. appl. water sci. doi:10.1007/s13201014-0232-4. amuzu at, 1975. water-treatment practices at barekese water utility in ghana. j. am. water work. assoc. 67:322–323. asante ka, quarcoopome t, amevenku fyk, 2008. water quality of the weija reservoir after 28 years of impoundment. west african j. appl. ecol. 13:125-131. ansa-asare do, asong-asante k, 1998. comparative study of the nutrient status of two reservoirs in southeast ghana. lakes reserv. res. manag. 3:205-217. annadotter h, cronberg g, lawton l, hansson hb, gothe u, skulberg om, 2001. an extensive outbreak of gastroenteritis associated with the toxic cyanobacterium planktothrix agardhii (oscillatorialis, no nco mm er cia l u se on ly g.n.d. addico et al.104 cyanophyceae) in scania, south sweden, p. 200-208. in: i. chorus (ed.) cyanotoxins occurrence, causes, consequences. springer-verlag, berlin. berger c, ba n, gugger m, bouvy m, rusconi f, coute a, troussellier m, bernard c, 2006. seasonal dynamics and toxicity of cylindrospermopsis raciborskii in lake guiers (senegal, west africa). fems microbiol ecol. 57:355-366. bernard c, ballot a, thomazeau s, maloufi s, furey a, mankiewicz-boczek j, pawlik-skowronska b, capelli c, salmaso n, 2017. appendix 2. cyanobacteria associated with the production of cyanotoxins, p. 503527. in: j. meriluoto, l. spoof and g.a. codd (eds.), handbook on cyanobacterial monitoring and cyanotoxin analysis. j. wiley & sons, chichester. bláha l, babica p, maršálek b, 2009. toxins produced in cyanobacteria water blooms. interdisc. toxicol. 2:36-41. bourke atc, hawes rb, nelson a, stallman nd, 1983. an outbreak of hepato-enteritis (the palm island mystery disease) possibly caused by algal intoxication. toxicon 21:45-48. byth s, 1980. palm island mystery disease. med. j. aust. 2:40-42. chen j, xie p, li l, 2009. first identification of the hepatoxic microsystins in the serum of a chronically exposed human population together with indication of hepatocellular damage. toxicol. sci. 108:81-89. chia ma, kwaghe mj, 2015. microcystins contamination of surface water supply sources in zaria-nigeria. env. monit. assess. 187:606. chia ma, abolude ds, ladan z, akanbi o, kalaboms a, 2009a. the presence of microcystins in aquatic ecosystems in northern nigeria: zaria as a case study. res. j. environ. toxicol. 3:170-178. chia ma, oniye sj, ladan zl, pila ae, inekwe vu, mmerole ju, 2009b. a survey for the presence of s in aquaculture ponds in zaria, northern-nigeria: possible health implications. african j. biotechnol. 8:6282-6289. fastner j, flieger i, neumann u, 1998. optimised extraction of microcystins from field samplesa comparison of different solvents and procedures. wat. res. 32:3177-3181. feurstein d, holst k, fischer a, dietrich dr, 2009. oatp-associated uptake and toxicity of microcystins in primary murine whole brain cells. toxicol. appl. pharmacol. 234:3372-3378. feurstein d, kleinteich k, heussner ah, stemmer k, dietrich dr, 2010. investigation of microcystin congener-dependednt uptake into primary murine neurons. environ. health perspect. 118:1370-1375. fleming le, rivero c, williams c, bean ja, shea ka, stinn j, 2002. blue green algal (cyanobacterial) toxins, surface drinking water, and liver cancer in florida. harmful algae 1:157-168. grosse y, baan r, kurt straif k, secretan b, el ghissassi f, cogliano v, 2006. carcinogenicity of nitrate, nitrite and cyanobacterial peptide toxins. lancet oncol. 7:628-629. haande s, ballot a, rohrlack t, fastner j, wiedner c, edvardsen b, 2007. diversity of microcystis aeruginosa isolates (chroococcales, cyanobacteria) from eastafrican water bodies. arch. microbiol. 188:15-25. harding wr, downing tg, van ginkel ce, moolma pma, 2009. an overview of cyanobacterial research and management in south africa post-2000. water sa. 35:479-484. harke mj, steffen mm, gobler cj, otten tg, wilhelm sw, wood sa, paerl hw, 2016. a review of the global ecology, genomics, and biogeography of the toxic cyanobacterium, microcystis spp. harmful algae 54:4-20. harada k, kondo f, lawton l, 1999. laboratory analysis of cyanotoxin, p. 369-399. in: ichorus and j. bartram (eds.) toxic cyanobacteria in water: a guide to their public health consequences, monitoring and management. e&fn spon, london. hoeger sj, hitzfeld b, dietrich d, 2005. occurrence and elimination of cyanobacterial toxins in drinking water treatment plants. toxicol. appl. pharmacol. 203:231-242. ho l, meyn t, keegan a, hoefel d, saint cp, and newcombe g, 2006. bacterial degradation of microcystin toxins within a biologically active sand filter. wat. res. 40:768-774. hrudey s, burch m, drikas m, gregory r, 1999. remedial measures, p. 275-312. in: ichorus and j. bartram (eds.) toxic cyanobacteria in water: a guide to their public health consequences, monitoring and management. e&fn spon, london. johansson s, olsson m, 1998. investigations of toxic algal blooms in algal blooms in a drinking water reservoir of harare, zimbabwe. msc. thesis, university of lunds, sweden. jones gj, orr pt, 1994. release and degradation of microcystin following algicide treatment of a microcystis aeruginosa bloom in a recreational lake, as determined by hplc and protein phosphatase inhibition assay. wat. res. 28:871-876 komárek j, anagnostidis k, 1999. [cyanoprokaryota. 1. teil: chroococcales], p. 1-548. in: h. ettl, g. gardner, h. heynig and d. mollenheuer (eds.), [süsswasserflora von mitteleurope].[book in german]. gustav fischer verlag. komárek j, anagnostidis k, 2005. [süßwasserflora von mitteleuropa, bd. 19/2: cyanoprokaryota].[book in german]. springer spektrum, heidelberg: 759 pp. kumasi tc, obiri-danso k, ephraim jh, 2011. microbial quality of water in barekese reservoir and feeder streams in ghana. lakes reserv. res. manag. 16:49-60. no nco mm er cia l u se on ly microcystins in drinking water in ghana 105 li y, chen ja, zhao q, pu c, qui z, zhang r, shu w, 2011. a cross-sectional investigation of chronic exposure to microcystin in relationship to childhood liver damage in the three gorges resources region, china, environ. health perspect. 119:1483-1488. manganelli m, scardala s, stefanelli m, palazzo f, funari e, vichi s, buratti fm, testai e, 2012. emerging health issues of cyanobacterial blooms. ann. ist. super. sanità. 48:415-428. mcelhiney j, lawton la, 2005. detection of the cyanobacterial hepatotoxins microcystins. toxicol. appl. pharm. 203:219-230. merel s, clement m, thomas o, 2010. state of the art on cyanotoxins in water and their behaviour towards chlorine. toxicon 55:677-691. merel s, walker d, chicana r, snyder s, baurés e, thomas o, 2013. state of knowledge and concerns on cyanobacterial blooms and cyanotoxins. environ. int. 59:303-327. mhlanga l, day j, cronberg g, chimbari m, siziba n, annadotter h, 2006. cyanobacteria and cyanotoxins in the source water from lake chivero, harare, zimbabwe, and the presence of cyanotoxins in drinking water. afr. j. aquat. sci. 31:165-173. mohamed za, carmchael ww, 2000. seasonal variation in microcystin levels of river nile water at sohag city, egypt. internat. j. limnol. 36:227-234. mowe mad, mitrovix sm, lim rp, furey a, yeo dcj, 2014. tropical cyanobacterial blooms: a review of prevalence, problem taxa, toxins and influencing environmental factors. j. limnol. 74:1005. myers jl, sekar r, richardson ll, 2007. molecular detection and ecological significance of the cyanobacterial genera geitlerinema and leptolyngbya in black band disease of corals. appl. environ. microbiol. 73:5173-5182. nasri ab, bouaicha n, fastner j, 2004. first report of a microcystin-containing bloom of the cyanobacteria microcystis spp. in lake oubeira, eastern algeria. arch environ. contam. toxicol. 46:197-202. ndebele mr, magadza chd, 2006. the occurrence of microcystin-lr in lake chivero, zimbabwe. lakes reserv. manage. 11:57-62 ndlela ll, oberholster pj, van wyk jh, cheng ph, 2016. an overview of cyanobacterial bloom occurrences and research in africa over the last decade. harmful algae 60:11-26. odokuma lo, isirima jc, 2007. distribution of cyanotoxins in aquatic environments in the niger delta. afr. j. biotechnol. 6:2375-2385. o’neil jm, davis tw, burford ma, gobler cj, 2012. the rise of harmful cyanobacteria blooms: the potential roles of eutrophication and climate change. harmful algae 14:313-334. pietsch j, bornmann k, schmidt w, 2002. relevance of intraand extracellular cyanotoxins for drinking water treatment. actahydrochim. hydrobiol. 30:7-15. quarcoopome t, amevenku f, ofori-danson p, 2011. changes in the fish community of the kpong headpond, lower volta river, ghana after 25 years of impoundment. rev. biol. trop. 59:1685-1696. stoler j, weeks j, fink, g, 2012. sachet drinking water in ghana’s accra-tema metropolitan area: past, present, and future. j. water sanit. hyg. dev. 2: 223-240. svircev z, krstic s, miladinov-mikov m, baltic v, vidovic m, 2009. freshwater cyanobacterial blooms and primary liver cancer epidemiological studies in serbia. j. environ. sci. health c environ. carcinog. 27:36-55. svircev z, baltic v, ganta m, jukovic m, stojanovic d, baltic m, 2010. molecular aspects of microcystin-induced hepatotoxicity and hepatocarcinogenesis. j. environ. sci. heal. part c-environmental carcinog. ecotoxicol. rev. 28:39-59. svircev z, drobac d, tokodi n, vidovic m, simevnovic j, mikov mm, baltic v, 2013. epidemiology of primary liver cancer in serbia and possible connection with cyanobacterial blooms. j. environ. sci. health c environ. carcinog. ecotoxicol. rev. 31:181-200. svircev z, drobac d, tokodi n, luzanin z, munjas zl, mikolin b, vuleta d, meriluoto j, 2014. epidemiology of cancers in serbia and possible connection with cyanobacterial blooms. j. environ. sci. health c environ. carcinog. ecotoxicol. rev. 32:319-337. ueno y, nagata s, tsutsumi t, hasegawa a, watanabe m, park hd, chen g, yu sz, 1996. detection of microcystins, a blue-green algal hepatotoxin, in drinking water sampled in haimen and fusui, endemic areas of primary liver cancer in china, by highly sensitive immunoassay. carcinogenesis 17:1317-1321. united nation statistical division, 2014. demographic yearbook. available from: https://unstats.un.org/unsd/ demographic/products/dyb/dyb2014.htm valdor r, aboal m, 2007. effects of living cyanobacteria, cyanobacterial extracts and pure microcystins on growth and ultrastructure of microalgae and bacteria. toxicon 49:769-779. van apeldoorn me, van egmond hp, speijers gja, bakker gji, 2007. toxins of cyanobacteria. mol. nutr. food res. 51:7-60. westrick ja, szlag dc, southwell bj, sinclair j, 2010. a review of cyanobacteria and cyanotoxins removal/inactivation in drinking water treatment. anal. bioanal. chem. 397:1705-1714. who, 1998. guidelines for drinking water quality cyanobacterial toxins: microcystin-lr world health organisation, who, geneva. no nco mm er cia l u se on ly g.n.d. addico et al.106 wood r, 2016. acute animal and human poisonings from cyanotoxin exposure a review of the literature. environ. int. 91:276-282. yu sz, 1989. drinking water and primary liver cancer, pp. 30-37. in: z.y. tang, m.c. wu and s.s. xia (eds.) primary liver cancer. springer, berlin. yu sz, 1995. primary prevention of hepatocellular carcinoma. j. gastroenterol. hepatol. 10:164-168. zhang f, lee j, liang s, shum ck, 2015. cyanobacteria blooms and non-alcoholic liver disease: evidence from a county level ecological study in the united states. environ. health 14:41. zhou l, yu h, chen k, 2002. relationship between microcystin in drinking water and colorectal cancer. biomed. environ. sci. 15:166-171. zilberg b, 1966. gastroenteritis in salisbury. european children a five year study. cent. afr. j. med. 12: 164-168. no nco mm er cia l u se on ly layout 1 introduction the need of systematic scientifically based monitoring programs is central for the management of the marine environment, which is facing salient alterations due to increasing and cumulative human impacts and the changing global climate (bax et al., 2018; miloslavich et al., 2018; mcquatters-gollop et al., 2019). the establishment and implementation of integrated, harmonized and extended monitoring systems and infrastructures have been also recognized as crucial for the achievement of the united nation sustainable development goals (sdgs), especially the un sdg 14 “life below water” (biermann et al., 2017) and the effective protection and restoration of the marine environment (mack et al., 2020). stemming from these emerging needs, marine observing systems (mos) have spread globally as integrated observing and experimenting infrastructures able to monitor oceanographic variables and to collect high-resolution data in selected marine regions to assess their state both in coastal and offshore areas (crise et al., 2018; rayner et al., 2019). mos, although essential for generating knowledge and supporting evidence-based decisions (goos, 2019), do not provide a truly holistic view of the marine ecosystems, since they mostly address the oceanographic processes and barely include the ecological ones. marine ecological observatories (meos) represent a further advancement of the mos perspective, since they broaden the spectrum of observing actions to embrace ecological research and monitoring (carr et al., 2011; benedetti cecchi et al., 2018; manea et al., 2020), recognizing ecological connectivity as one of the main driving forces of marine ecosystems’ functioning. ecological connectivity represents “the degree to which landscapes and seascapes allow species to move freely and ecological processes to function unimpeded” (unep, 2019). due to the intrinsic dynamism that characterizes the marine realm, ecological connectivity is a crucial element, embracing the complex interconnections among the natural processes, the species, their life cycles and the environment (carr et al., 2003 and 2011; maxwell et al., 2015; manea et al., 2019). within this broad perspective, coherent and operative meos are recommended at the european and global level (benedetti cecchi et al., 2018; carr et al., 2011; duffy et al., 2013; muelbert et al., 2019). meos can deliver great benefits to enhance both the understanding of the marine ecosystems and the assessment of their state and changes over time and space, and the monitoring of the sea for protection and conservation purposes. this can be partly done by integrating data derived from monitoring practices routinely carried out to comply with policy requirements, from long-term ecological research (e.g., lter data, acri et al., 2020), from ecological research projects and mpas monitoring plans, and from enlarging the set of variables measured by observing infrastructures to include biodiversity (e.g., species distribution data through video imagery, aguzzi et al., 2020). among the many attributes and uses meos should inreview a transnational marine ecological observatory in the adriatic sea to harmonize a fragmented approach to monitoring and conservation elisabetta manea,1 caterina bergami,2* lucia bongiorni,1 lucilla capotondi,2 elisabeth de maio,1 alessandro oggioni,3 alessandra pugnetti1 1institute of marine sciences, national research council, ismar-cnr, arsenale, tesa 104, castello 2737/f, 30122 venice; 2institute of marine sciences, national research council, ismar-cnr, via piero gobetti 101, 40129 bologna; 3institute for electromagnetic sensing of the environment, national research council, irea-cnr, via bassini 15, 20133 milan, italy abstract the design and establishment of marine ecological observatories (meos) are fostered at the european level. meos should adopt a holistic view, integrating and harmonizing long-term oceanographic and ecological research and monitoring, and increasing conservation strategies effectiveness according to the ecological connectivity concept. the data and knowledge collected and made available through meos should inform policies dealing with conservation and management of the marine environment. we present and discuss these issues in the adriatic sea context, where the transnational ecological observatory “ecoads” is under development in the framework of the interreg italy-croatia project ecoss (observing system in the adriatic sea: oceanographic observations for biodiversity), which aims to support natura 2000 network implementation and cross-border coordination of multiple monitoring initiatives. we analyse the main eu directives that deal with marine conservation, notably the habitats and birds directives, the water framework directive, and the marine strategy framework directive, as they are the primary instruments that can guide the development of the observatory while strengthening cooperation at the basin scale. we bring out the synergies and discrepancies among these legal instruments and build on them ecoads as a monitoring platform that may respond and contribute to their requirements, boosting the synergies and overcoming the weaknesses. finally, we provide some hints for the further development of this transnational meo as a collector of the existing monitoring efforts aimed at harmonizing their approaches and incorporating the ecological connectivity to foster an ecosystem-based approach to conservation management. no nco mm er cia l u se on ly e. manea et al.12 corporate (benedetti cecchi et al., 2018; manea et al., 2020), in this paper we focus on the role they should cover to inform marine conservation strategies and to feed european union (eu) policies and strategies that deal with coastal and marine management. different jurisdictional instruments are in force in europe to these purposes, which could take advantages from and contribute to meos, in particular the habitats and birds directives (eec 1992 and ec 2009; hd and bd, respectively), the water framework directive (ec 2000; wfd), and the marine strategy framework directive (ec 2008; msfd). the european commission developed in 2017 an action plan (com(2017) 198 final) to improve the implementation of the nature directives (hd and bd, respectively) and the related natura 2000 (n2k) network, which is the main conservation instrument put in place to accelerate the achievement of eu 2020 conservation goals. among the diverse priorities and related actions to be taken, the plan highlights the need of enhanced and efficient monitoring, as well as of completing the n2k network by filling knowledge gaps to feed the required conservation measures. the wfd and the msfd were also established to provide legal instruments able to support the conservation and sustainable management of the coastal and marine environment (kallis and butler, 2001; borja et al., 2010; long 2011), and they strongly rely on monitoring programs proposing operational frameworks and approaches. commonalities and synergies among the above-mentioned eu directives are expected, since they all tend to support the achievement of declared conservation goals, and they have been partly analysed from the policy and management perspectives (rouillard et al., 2018). meos, by harmonizing different monitoring frameworks and by collecting and transferring oceanographic and ecological data, can be key knowledge sharing platforms to support the directives and their monitoring plans, enhancing the benefits they could mutually provide to achieve environmental goals. we present and discuss these issues in the context of the adriatic sea, where the ecological observatory “ecoads” is under design and development in the framework of the interreg italy-croatia project ecoss (observing system in the adriatic sea: oceanographic observations for biodiversity; https://www.italycroatia.eu/web/ecoss). indeed, ecoads pilot case will integrate the existing ecological and oceanographic research monitoring with n2k conservation strategies, to boost nature directives and the implementation of n2k network in the marine area of interest for italy and croatia. here we analyse the above-mentioned eu directives, critical instruments able to guide the development of the observatory while supporting the strengthening of transnational cooperation at the basin scale. we bring out the synergies and discrepancies among these legal instruments and build on them ecoads as a monitoring platform that may respond and contribute to their requirements, boosting the synergies and overcoming the weaknesses. the main aim is to compare their monitoring approaches to start harmonizing a set of oceanographic and ecological variables to be monitored at the adequate spatial and temporal scales, able to assess the environmental state and the n2k sites’ effectiveness, with the additional intent to support the globally recognized essential biodiversity variables (ebvs) and essential ocean variables (eovs) frameworks (pereira et al., 2013; mullerkarger et al., 2018). we provide some hints regarding the on-going and future activities that are accompanying the development of this transnational meo. the proposed approach will incorporate ecological connectivity across spatial and temporal scales to inform the future expansion of n2k network and to support the application of an ecosystem-based approach to conservation (unep, 2011). ecoads, the marine ecological observatory (meo) of the adriatic sea the need of a meo in the adriatic ecoads focuses on the area covered by the interreg italy-croatia program and the adjacent adriatic sea, which hosts an outstanding biodiversity and unique environmental conditions, thus delivering multiple ecosystem services and benefits. these include provisioning of food (the adriatic is among the main fish areas of the mediterranean sea, scarcella et al., 2014) and material, carbon sequestration, water quality regulation and coastal protection, as well as opportunities of leisure and spiritual benefits to people (depellegrin et al., 2017; cossarini et al., 2015). however, this marine area is barely covered by marine protected areas (mpas; bastari et al., 2016, manea et al., 2019). furthermore, it is strongly affected by a huge variety of human activities that have found fertile ground on which to develop, among which fishery, aquaculture, maritime traffic, and land-based activities that affect nutrients and contaminants load reaching the marine environment (farella et al., 2020; milliman et al., 2016). for this reason, the adriatic sea has been recognized as one of the most impacted area of the mediterranean sea (micheli et al., 2013). the implementation of n2k network in the basin is on track with some delay, and it mostly lacks suitable and fulfilled management plans and adequate monitoring programs (claudet et al., 2020). nevertheless, several fixed-point observing systems (i.e., pylons, buoys, tide gauges, oceanographic platforms; fig. 1) are active (ravaioli et al., 2016; šepić and vilibic, 2011; šepić et al., 2017), providing multidisciplinary data and information automatically and continuously no nco mm er cia l u se on ly a transnational marine ecological observatory in the adriatic sea 13 fig.1. natura 2000 network, national marine protected areas, sites and fixed point observing systems in the adriatic sea. part of these observing systems is managed by ecoss project’s partners. data sources: mapamed, 2017 (https://www.rac-spa.org/mapamed); veneto region web portal (https://bur.regione.veneto.it/burvservices/pubblica/dettagliodgr.aspx?id=426153); emilia romagna region web portal (https://bur.regione.emilia-romagna.it/dettaglio-inserzione?i=dc40e3d31e384e718b81c23cad7946e9); ravaioli et al., 2016; meteo-tsunami network (http://jadran.izor.hr/hazadr/index_eng.htm); tide gauges network (https://www.hhi.hr/en; https://www.mareografico.it); lter-italy website (http://www.lteritalia.it/). no nco mm er cia l u se on ly e. manea et al.14 over time. different monitoring initiatives are also ongoing, as well as specific programs, such as the italian longterm ecological research network (lter-italy), but at the moment none of these initiatives has a cross-border nature shared between italy and croatia. furthermore, only few of these are operatively supporting the monitoring observations linked to the fulfilment of the eu obligations, despite the wide and varied array of data collected, ranging from those related to the quality of transitional, coastal and marine waters, to the monitoring of target species and habitats (e.g., dolphins, sea turtles, seagrass meadows, coralligenous outcrops). these observing platforms and monitoring schemes operate at various scales, with different aims and maturity levels, lacking an adequate coordination among them, which should connect the local, the regional, up to the whole adriatic basin scale, in an integrated and coherent observatory. in this context, ecoads gives the opportunity to define and adopt an agreed conceptual framework, to harmonize existing environmental observations, monitoring schemes, and descriptive indicators (manea et al., 2020; fig. 2), thus contributing at providing a complete picture of the state of the marine environment, and at supporting transnational cooperation between italy and croatia in the adriatic sea. the directives in the design of ecoads the design phase of ecoads entails, as a first crucial step, an accurate comparative analysis among the main eu legal instruments focused on coastal and marine environment protection, namely the wfd, the msfd, the hd and bd. for the hb and bd, which act on both terrestrial and aquatic environments, the analysis considers fig. 2. schematic view of the essential attributes of ecoads to support monitoring programs harmonization. ecoads embraces all the observing systems and monitoring programs managed by italy and croatia in the adriatic sea, by integrating an open science approach and the ecological connectivity concept and providing a harmonized set of monitoring indicators to support transnational cooperation in conservation management. ecoads addresses the requirements of the main european directives by linking their monitoring programs to those already existing in the area covered by the interreg italy-croatia program, and to natura 2000 sites. no nco mm er cia l u se on ly a transnational marine ecological observatory in the adriatic sea 15 exclusively coastal and marine conservation objectives and targets. the primary aim of such analysis is to identify how ecoads could contribute to support the directives’ implementation and to overcome possible limitations for the effective monitoring of the aquatic environment. to achieve this, we ascertain the main directives’ synergies and commonalities, assessing and highlighting the opportunities that could emerge if synergies would be underpinned, and we individuate potential weaknesses and inconsistencies that should be overcome to support conservation strategies. firstly, a short review of the directives is conducted considering key features characterizing them using multiple knowledge sources, mainly coming from official documentation, such as commission directives, reports and guidelines. the key features are (tab. 1): i) conservation objectives and targets; ii) approaches adopted to foster conservation initiatives; iii) spatial application; iv) reporting period that guides the evaluation of conservation objectives and targets achievement and of conservation measures effectiveness; v) considerations related to the human uses, the human-derived pressures, and the provision of marine ecosystem services; vi) the criteria and indicators adopted to evaluate conservation objectives and targets achievement; and vii) the indications related to the monitoring strategies. the comparative analysis is firstly built to allow the explicit emergence of the synergies and weaknesses among the legal instruments. then, the analysis is zoomed in on the monitoring framework entailed within each directive. finally, the main elements deriving from the analysis are discussed in the context of the future development of ecoads. eu directives for marine conservation: description, comparison and monitoring framework in tab. 1 a comparison between the four directives is performed, hinging on the seven features described above. related insights on synergies, weaknesses and monitoring approaches are presented in the following sections. synergies among directives good quality of ecological and environmental state is central for all the directives as they express the need of establishing conservation measures. the hd and bd, by supporting the implementation of the n2k network, are more focused on the in situ protection (grodzińska-jurczak et al., 2012; evans, 2012), while the wfd and msfd address a broader scale and extend their conservation strategies also beyond the establishment of areas dedicated to conservation. although not specifically designated for conservation, the wfd shows some relevant connections with nature conservation issues. indeed, it mentions the protection of terrestrial ecosystems and wetlands directly depending on the aquatic ones, clearly referring to the management objectives of n2k sites. the msfd recognizes the important contribution that the establishment of mpas, including n2k sites, delivers to the achievement of good environmental status (gens). thus, the msfd promotes the designation of mpas with the aim of contributing, together with the n2k network, to the achievement of the conservation targets. if the hd and bd focus on the conservation of target species and habitats, the wfd and the msfd expand their action by including additionally functional aspects and considering the ecosystems as a conservation unit. these directives can be considered complementary to each other as together provide an overall normative context for conservation priority objectives at different spatial and ecological levels. the wfd and the msfd apply within different jurisdictional boundaries, the first being focused on internal and coastal waters (up to 1 nm), and the second extending from the coast to offshore areas. consequently, if appropriately coordinated, they cover the entire water territories supporting the implementation of n2k network also in offshore areas, not leaving gaps amidst aquatic domains. indeed, n2k sites are spread from the coast to the offshore, and the sites can include a mix of habitats and species being composed by terrestrial, freshwater and marine ecosystems, even crossing the land-sea interface (kati et al., 2015). all these legal instruments address the need to ensure a sustainable management of natural resources, by balancing the societal economic needs and the conservation priorities and highlighting the tight interconnections between nature and humans. explicitly or implicitly, the four directives consider the ecosystem-based approach (rouillard et al., 2018), they mention natural goods and services, recognizing the critical need of preserving and managing them. while the hd, the bd and the wfd only generically mention the human activities, the msfd makes an ad hoc analysis of the human-derived pressures and impacts that can negatively affect the marine environment, highlighting the need of addressing all of them through ecosystem-based management measures. taking advantage of the analysis made by the msfd, the n2k network managers can be better supported in setting adequate monitoring programs for an improved n2k sites management. finally, all the directives require monitoring practices to track their goals achievement. the reporting periods of the four directives mainly overlap, except for the wfd that reports every three years instead of six. the provided information entailed within all reporting periods potentially furnishes a complete and comprehensive knowledge framework to inform management and conservation strategies. no nco mm er cia l u se on ly e. manea et al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a> ? @b ;! "!# $% &' $% &! &$ '5 /$ 7!, %/ -0 c!: *: 57 $' %* &0 ! %& !'9 (! d %7!0 '$ '( !'* !( )* 7* 4% )$ 7" 0) %( &' %+ %) ") 57 '5 /$ 7!$ -( 85 $' (! 7( .( 70 ;!! "!< /* '( )' %* &! *+ !0 (7 () '( -! 0: () %( 0! $& -! 9$ ,% '$ '0 !%(& '%+ %( -! $0 !* +!: /% */ %'1 !+* /! :/ *' () '%* &; ! "!e 9( !: /( 0( /. $' %* &2 !6 $% &' (& $& )( !* /! /( 0' */ $' %* &! *+ !$ !0 5+ +% )% (& '!, %/ -0 c! -% .( /0 %'1 !$ &!$ /( $! *+ !9 $, %'$ '0 ;! "!# $% &' $% &! $& -! %6 :/ *. (! '9 (! 85 $7 %'1 !* +!' 9( ! $8 5$ '%) !( &. %/ *& 6 (& '!% &! '9 (! ? *6 6 5& %'1 ;!! "!= )9 %( .( 6 (& '!* +!f ** -! g )* 7* 4% )$ 7! @' $' 50 h! af g )@ b!* +!$ 77! d $' (/ 0! ,1 !i jk l2 !* /! ij ik ;! "!= )9 %( .( 6 (& '!* +!' 9( !f ** -! ? 9( 6 %) $7 ! @' $' (! af ? @b !* +!$ 77! d $' (/ 0! ,1 !i jk l2 !* /! ij ik ;! "!< /* '( )' %* &! $& -! )* &0 (/ .$ '%* &! *+ !'9 (! 6 $/ %& (! (& .% /* &6 (& '2! :/ *6 *' %* &! *+ ! '9 (! 05 0' $% &$ ,7 (! 50 (! *+ !'9 (! 0( $0 2!$ &! )* &0 (/ .$ '%* &! 6 $/ %& (! () *0 10 '( 6 0; !! "!= )9 %( .( 6 (& '!$ &!6 $% &' (& $& )( !* +! f ** -! g &. %/ *& 6 (& '$ 7!@ '$ '5 0m !af g &@ b! *+ !6 $/ %& (! d $' (/ 0! ,1 !i ji j; ! n %/ () '%. (0 o!g g = ? 2! ij kp ! !!" #5 44 '* (6# .* # -* &0 %' /( .!* &# "!@ :$ '%$ 7!$ &!6 $& $4 (6 (& '! 6 ($ 05 /( 0! 6 $% &7 1! '9 /* 54 9! '9 (! (0 '$ ,7 %0 96 (& '!* +!' 9( !g q ! () *7 *4 %) $7 !r $' 5/ $! ij jj !r (' d */ 3! *+ !: /* '( )' (!0 %'( 0! a@ :( )% $7 != /( $0 ! *+ !? *& 0( /. $' %* &! s! @= ? 0b ;!! "!< /* $) '%. (! $& -! /( $) '%. (t !, (1 *& -! )* &0 (/ .$ '%* &! 6 ($ 05 /( 02 !%' ! )* &0 %(/ 0! /( 0' */ $' %* &! $& -! /( )* .( /1 ! $) '%* &0 ;! "!e 9( 1! d */ 3! 6 */ (! $' !& $' %* &$ 7!7 (. (7 ! (. (& !'9 *5 49 !/( 4% *& $7 !) ** :( /$ '%* &! %0 !9 %4 97 1! /( )* 6 6 (& -( -; ! "!u %0 '!* +!0 :( )% (0 !$ &!9 $, %'$ '0 !'* ! :/ *' () '!$ &!6 *& %'* /!$ 7/ ($ -1 ! -( +% &( -; ! "!@ :$ '%$ 7!$ &!6 $& $4 (6 (& '!6 ($ 05 /( 0! 6 $% &7 1! '9 /* 54 9! '9 (! (0 '$ ,7 %0 96 (& '!* +! '9 (! g q !( )* 7* 4% )$ 7!r $' 5/ $! ij jj ! r (' d */ 3! *+ !: /* '( )' (!0 %'( 0! a@ :( )% $7 ! </ *' () '%* &! = /( $0 !"! @< = 0b ;! "!< /* $) '%. (! $& -! /( $) '%. (t !, (1 *& -! )* &0 (/ .$ '%* &! 6 ($ 05 /( 02 !%' !) *& 0% -( /0 ! /( 0' */ $' %* &! $& -! /( )* .( /1 !$ )' %* &0 ;! "!e 9( 1! d */ 3! 6 */ (! $' !& $' %* &$ 7!7 (. (7 ! (. (& !'9 *5 49 !/( 4% *& $7 !) ** :( /$ '%* &! %0 ! 9% 49 71 !/( )* 6 6 (& -( -; ! "!u %0 '!* +!, %/ -! 0: () %( 0! $& -! 9$ ,% '$ '0 !'* ! :/ *' () '!$ &!6 *& %'* /!$ 7/ ($ -1 !(+ %& (;! "!g )* 7* 4% )$ 7!0 '$ '( !%& )7 5%& 4! () *0 10 '( 6 ! +5 &) '%* &% &4 2!, (1 *& -! '9 (! :/ *' () '%* &! *+ ! 0% &4 7( !0 :( )% (0 !$ &!9 $, %'$ '0 ;!! "!n () *& 0' /5 )' %& 42 !0 '/ 5) '5 /$ 7!$ :: /* $) 9v 2! ,$ 0( -! *& !l !, %* 7* 4% )$ 7!8 5$ 7%' 1! (7 (6 (& '0 ! aw x g b!: 75 0! 91 -/ *6 */ :9 *7 *4 %) $7 !$ &! :9 10 %) *) 9( 6 %) $7 !8 5$ 7%' 1! (7 (6 (& '0 ;! "!< /* $) '%. (! $& -! /( $) '%. (t !, (1 *& -! )* &0 (/ .$ '%* &! 6 ($ 05 /( 02 !%' !) *& 0% -( /0 ! /( 0' */ $' %* &; ! "!y *7 %0 '%) 2!+ 5& )' %* &$ 7!$ :: /* $) 9; ! "!< /* $) '%. (! $& -! /( $) '%. (t !, (1 *& -! )* &0 (/ .$ '%* &! 6 ($ 05 /( 02 !%' !) *& 0% -( /0 ! /( 0' */ $' %* &! $& -! /( )* .( /1 !$ )' %* &0 ;! "!g z: 7%) %'7 1! () *0 10 '( 6 ", $0 (;! "!y %4 97 1! '/ $& 0& $' %* &$ 7!a &( (0! '* ! d */ 3! '9 /* 54 9! /( 4% *& $7 !: /* 4/ $6 6 (0 b;! "!? *& 0% -( /$ '%* &! *+ !0 *) %* "( )* &* 6 %) ! $0 :( )' 0; ! "![ '!( z: 7%) %'7 1! %& )7 5(0 !( )* 7* 4% )$ 7! )* && () '%. %'1 ;! "![ '!% &' (4 /$ '( 0! '9 (! :/ () $5 '%* &$ /1 ! $: :/ *$ )9 ;! n %/ () '%. (0 o!w */ \$ !i jj lo ! = 77$ &! %. #( )72 !i jj ]o ! w */ \$ !% .#( )72 !i jk jo !# (( ! %. #( )72 !i jj pb o!> = x !+% &$ 7! ij ki "j ^" i^ o! 9' ': 0t __( ); (5 /* :$ ;( 5_ (& .% /* &6 (& '_& $' 5/ (_ &$ '5 / $i jj j_ 6 $/ %& (_ -* )0 _> = x ` ij +% &$ 7` ij ij ki " j^ "i ^; :+! !!! "#8 4( .!( )# (4 4) !(. !* &# "!w /* $;! "!a !6 $% &! ,% *4 (* 4/ $: 9% )! /( 4% *& 0! d %'9 !/( :/ (0 (& '$ '%. (! 9$ ,% '$ '0 !$ &! 0: () %( 0; ! "!e /$ &0 %'% *& $7 2!) *$ 0' $7 !$ &! '( // %'* /% $7 !d $' (/ 02 !%& )7 5%& 4! g g b ! d 9( /( !() 7$ /( -; ! "!9 &# 0! .: !: /* '( )' %* &; ! "!w /* $;! "!g 5/ *: ($ &! '( // %'* /1 !$ &!& *' ! ,% *4 (* 4/ $: 9% )! /( 4% *& 0; ! "!e /$ &0 %'% *& $7 2!) *$ 0' $7 !$ &!'( // %'* /% $7 ! d $' (/ 02 !%& )7 5%& 4! g g b !d 9( /( ! -( )7 $/ (;! "!c %. (/ 02 !7$ 3( 02 !'/ $& 0% '%* &$ 7!$ &!) *$ 0' $7 ! d $' (/ 0! 5: !'* !k !& 6 !+/ *6 !'9 (! )* $0 '7% &( ;! "!c %. (/ !, $0 %& !%0 '/ %) '0 !$ 0! 6 $& $4 (6 (& '! 5& %'0 !+* /!c %. (/ !w $0 %& 0! # $& $4 (6 (& '!< 7$ &0 ;! "!w /* $;! "!# $/ %& (! /( 4% *& 0! '9 $' !%& )7 5(! 05 ," /( 4% *& 0; ! "!? *$ 0' $7 !$ &!'( // %'* /% $7 !d $' (/ 02 ! %& )7 5%& 4! g g b !d 9( /( !() 7$ /( -; ! n %/ () '%. (0 ! !/ "#; %4 *' .!& 2# 4% '! *1 # "!c (: */ '%& 4! (. (/ 1! ]! 1( $/ 0! *& !'9 (! %6 :7 (6 (& '$ '%* &! *+ !& $' %* &$ 7! :/ *. %0 %* &0 !'$ 3( &! 5& -( /!' 9( ! -% /( )' %. (! $& -! '9 (% /!( ++ () '%. (& (0 0; ! "!r (( -! *+ !6 *& %'* /% &4 !'* ! d5 &(/ '$ 3( !0 5/ .( %77 $& )( !* +!' 9( ! )* &0 (/ .$ '%* &! 0' $' 50 !* +!' 9( !& $' 5/ $7 ! 9$ ,% '$ '0 !$ &!0 :( )% (0 e; ! "!f /% 4% &$ 771 !/( :* /' %& 4! (. (/ 1! g2 !7$ '( /!] ! 1( $/ 0! *& !'9 (! %6 :7 (6 (& '$ '%* &! *+ ! &$ '%* &$ 7!: /* .% 0% *& 0! '$ 3( &! 5& -( /!' 9( ! -% /( )' %. (; ! ! "!# *& %'* /% &4 !$ &!/( :* /' %& 4! (. (/ 1! g! 1( $/ 0; ! "!# *& %'* /% &4 !$ &!/( :* /' %& 4! (. (/ 1! ]! 1( $/ 0; ! >= x !+% &$ 7!i jk i" j^ "i ^o ! n %/ () '%. (0 o!w */ \$ !i jj l! ! ta b. 1 . c om pa ri so n am on g th e m os t r el ev an t j ur is di ct io na l i ns tr um en ts c on ce iv ed fo r t he c on se rv at io n of th e aq ua tic e nv ir on m en t b as ed o n se le ct ed fe at ur es . c om pa ri so n fe at ur es h d (9 2/ 43 /e e c ) b d (2 00 9/ 15 7/ e c ) w fd (2 00 0/ 60 /e c ) m sf d (2 00 8) r ef er en ce s to b e co nt in ue d on n ex t p ag e no nco mm er cia l u se on ly a transnational marine ecological observatory in the adriatic sea 17 /" #< := (& #( -. !/ !.! %0 ># 1% '! /% 1# 4' %0 0: '% 0> # (& 1# %*0 ?0 .% = # 0% '/ !%0 #@a 8" # "!# $% &7 1! 95 &' %& 42 !%7 7( 4$ 7!3 %77 %& 42 ! '/ $: :% &4 !$ &!'/ $(! *+ !0 :( )% (0 !!! "![ &! %'0 !* ,\ () '%. (0 !'9 (! -% /( )' %. (! (& '$ %70 !) *& 0% -( /$ '%* &0 !* +! () *& *6 %) 2!0 *) %$ 72! )5 7'5 /$ 7!$ &! /( 4% *& $7 !$ 0: () '0 !/* *' (!%& !'9 (! 50 (! *+ !& $' 5/ $7 !/( 0* 5/ )( 02 !6 $3 %& 4! $! )* &' /% ,5 '%* &! '* !'9 (! 05 0' $% &$ ,7 (! -( .( 7* :6 (& ';! r *' !( z: 7%) %'! /( +( /( &) (! '* !g @; ! "!# $% &7 1! 95 &' %& 42 !%7 7( 4$ 7!3 %77 %& 42 ! '/ $: :% &4 !$ &!'/ $(! *+ !, %/ -0 ! "!u *& 4" '( /6 !: /* '( )' %* &! $& -! 6 $& $4 (6 (& '!* +!& $' 5/ $7 !/( 0* 5/ )( 0! $0 ! $& !%& '( 4/ $7 !: $/ '!* +!' 9( !9 (/ %'$ 4( !* +!' 9( ! :( *: 7( 0! *+ !g 5/ *: (; ! "!? *& '/ *7 !& $' 5/ $7 !/( 0* 5/ )( 0! $& -! 4* .( /& 0! '9 (% /!5 0( !, $0 (!* &! '9 (! 6 ($ 05 /( 0! &( )( 00 $/ 1! +* /!' 9( ! 6 $% &' (& $& )( !$ &!$ -\ 50 '6 (& '!* +!' 9( ! &$ '5 /$ 7!, $7 $& )( 0! ,( 'd (( &! 0: () %( 0! $0 ! +$ /!$ 0! %0 !/( $0 *& $, 71 !: *0 0% ,7 (; !r *' ! (z :7 %) %'! /( +( /( &) (! '* !g @; ! "!# (6 ,( /!@ '$ '( 0! 09 *5 7!( &0 5/ (! $! /( .% (d ! *+ !'9 (! %6 :$ )' !* +!9 56 $& !$ )' %. %'1 !* &! '9 (! 0' $' 50 !* +!0 5/ +$ )( !d $' (/ 0! $& -! *& ! 4/ *5 &d $' (/ ! "!# $% &7 1! +* )5 0( 0! *& !: *7 75 '%* &! 0* 5/ )( 0! $& -! '* !/( -5 )( !'9 (! -% 0) 9$ /4 (! $& -! (6 %0 0% *& !* +! :* 775 '$ &' 0! $& -! 9$ h$ /*5 0! 05 ,0 '$ &) (0 ! "!? 7( $/ !/( +( /( &) (! '* !d $' (/ !: /* .% 0% *& %& 4! $0 ! +5 &$6 (& '$ 7!0 (/ .% )( !'* !9 56 $& !!! "!e 9( !0 '/ $' (4 1! $-/ (0 0( 0! $7 7!9 56 $& ! $) '%. %'% (0 !'9 $' !9 $. (! $& !%6 :$ )' !* &! '9 (! 6 $/ %& (! (& .% /* &6 (& ';! "!y 56 $& "(/ %. (!: /( 00 5/ (0 !$ &! %6 :$ )' 0! '* !6 *& %'* /!$ &!6 $& $4 (! 0: () %+ %( -! $& -! -( 0) /% ,( -! a= && (z ![[ [b ;! "!c (+ (/ (& )( !'* !'9 (! $*: '%* &! *+ !$ &! () *0 10 '( 6 ", $0 (!$ :: /* $) 9! $& -! '* ! '9 (! %6 :* /' $& )( !* +!' 9( !6 $/ %& (! (& .% /* &6 (& '!+ */ !'9 (! 0( /. %) (0 !$ &! ,( &( +% '0 !%' !(7 %. (/ 0! '* !9 56 $& 0; ! n %/ () '%. (0 o! 9' ': 0t __( ); (5 /* :$ ;( 5_ (& .% /* &6 (& '_& $' 5/ (_ 7( 4% 07 $' %* &_ ,% /0%/ () '%. (_ %& (z i( &; 9' 6 o! g q !i jk ^_ pj p! !! /! "#b '! .% '! (# (& 1# c %' 3* '= (& -% # !& 1! -( .* '# "!? /% '( /% $! +* /!0 %'( !0 (7 () '%* &! ,* '9 ! +* /!$ !4 %. (& !& $' 5/ $7 !9 $, %'$ '!$ &! 0: () %( 02 != && (z ![[ [;! "!= )9 %( .( 6 (& '!* /!/ (0 '* /$ '%* &! *+ ! '9 (! >$ .* 5/ $, 7( !? *& 0( /. $' %* &! @' $' 50 !a> ? @2 !> $. *5 /$ ,7 (2 ! 5& +$ .* 5/ $, 7( !"! %& $(8 5$ '( 2! 5& +$ .* 5/ $, 7( !"! ,$ -2 !5 &3 &* d &b !$ &! $0 0* )% $' (!) /% '( /% $! +* /!, *' 9! 9$ ,% '$ '0 !$ &!0 :( )% (0 ;! "!r *' !%& -% )$ '( -! (z :7 %) %'7 1! %& !'9 (! -% /( )' %. (; !> */ !0 (7 () '%* &! *+ !r ik !0 %'( 0! +* /!, %/ -0 c!: /* '( )' %* &2 !'9 (! )/ %'( /% $! $/ (! '9 (! 0$ 6 (! %& -% )$ '( -! %& !'9 (! y n ;! "!= )9 %( .( 6 (& '!* /!/ (0 '* /$ '%* &! *+ !'9 (! >$ .* 5/ $, 7( !? *& 0( /. $' %* &! @' $' 50 !a> ? @2 ! >$ .* 5/ $, 7( 2!5 &+ $. *5 /$ ,7 (! "! %& $(8 5$ '( 2!5 &+ $. *5 /$ ,7 (! "!, $2! 5& 3& *d &b !$ &!$ 00 *) %$ '( -! )/ %'( /% $! +* /! ,* '9 !9 $, %'$ '0 !$ &!0 :( )% (0 ;! "!f ** -! g )* 7* 4% )$ 7!@ '$ '5 0! af g )@ b;! ! "!f ** -! ? 9( 6 %) $7 !@ '$ '( !* +!$ 77! d $' (/ 0; !! "!y %4 92 !4 ** -2 !6 *(/ $' (2 !: ** /2! ,$ -! 85 $7 %'1 ! *+ !'9 (! () *7 *4 %) $7 !0 '$ '5 0! -( +% &( -! ,1 ! &* /6 $' %. (; ! "!f ** -! g &. %/ *& 6 (& '$ 7!@ '$ '5 0! af g &@ b;! "!k k! -( 0) /% :' */ 0; ! "!f ** -! */ !& *' !4 ** -; ! n %/ () '%. (0 ! /! !" #9& 1! -( .!* &# 3* '# = *& !.* '! &2 # "!c (8 5% /( -! 6 *& %'* /% &4 !, 5' !& *' ! (z :7 %) %'! %& -% )$ '%* &; ! "!= -* :' %* &! *+ !4 (' (= %. %' 0! +* /!0 %'( ! 0( 7( )' %* &! $& -! 6 *& %'* /% &4 ! "!c (8 5% /( -! 6 *& %'* /% &4 !, 5' !& *' !( z: 7%) %'! %& -% )$ '%* &; ! "!= -* :' %* &! *+ !4 (' (= %. %' 0! +* /!0 %'( ! 0( 7( )' %* &! $& -! 6 *& %'* /% &4 ! "!n (' $% 7( -! 6 *& %'* /% &4 !/( 85 %/ (6 (& '0 !a( ;4 ;2! '1 :( 0! *+ !6 *& %'* /% &4 2!6 *& %'* /% &4 ! +/ (8 5( &) 1b ;! "!= -* :' %* &! *+ !d :( )!. ?# %) %= %& .0 !'* !, (! 6 *& %'* /( -! "!n (' $% 7( -! 6 *& %'* /% &4 !/( 85 %/ (6 (& '0 ;! "!g 0' $, 7%0 9( -! )/ %'( /% $! $& -! 6 (' 9* -* 7* 4% )$ 7!0 '$ &$/ -0 !* &! f ** -! g &. %/ *& 6 (& '$ 7!@ '$ '5 0! af g &@ b!$ &! 0: () %+ %) $' %* &0 !$ &!0 '$ &$/ -% 0( -! 6 (' 9* -0 !+* /!6 *& %'* /% &4 !$ &! $0 0( 00 6 (& ';! "!= -* :' %* &! *+ !'! .% '! (! 7%& 3( -! '* !'9 (! kk !(0 )/ %: '* /0 !'* !, (! 6 *& %'* /( -! n %/ () '%. (0 o!g q ! ij k^ _p jp ! y n 2!9 $, %'$ '0 !%/ () '%. (o !w n o!, %/ -0 !%/ () '%. (o !l >n o!d $' (/ !+/ $6 (d */ 3! n %/ () '%. (o !# @> n 2!6 $/ %& (! 0' /$ '( 41 !+/ $6 (d */ 3! -% /( )' %. (; !h f g )@ !%0 !(+ %& (!$ 0! '9 (! .$ 75 (0 !* +!' 9( !, %* 7* 4% )$ 7!8 5$ 7%' 1! (7 (6 (& '0 !+* /!' 9( !0 5/ +$ )( !d $' (/ !, *1! '1 :( !d 9% )9 !0 9* d !7* d !7( .( 70 !* +! -% 0' */ '%* &! /( 05 7'% &4 !+ /* 6 !9 56 $& !$ )' %. %'1 2!, 5' !(. %$ '( !* &7 1! 07 %4 9' 71 !+ /* 6 !'9 *0 (! &* /6 $7 71 !$ 00 *) %$ '( -! d %'9 !'9 (! 05 /+ $) (! d $' (/ !, *1! '1 :( !5 &(/ !5 &%0 '5 /, (!) *& -% '%* &0 ;!m f g &@ !%0 !(+ %& (!$ 0! '9 (! (& .% /* &6 (& '$ 7!0 '$ '5 0!* +!6 $/ %& (! d $' (/ 0!d 9( /( !'9 (0 (! :/ *. %(! () *7 *4 %) $7 71 !%. (/ 0( !$ &!1& $6 %) !* )( $& 0!$ &!0( $0 !d 9% )9 !$ /( !) 7( $& 2!9 ($ 7'9 1! $& -! :/ *5) '%. (! d %'9 %& !'9 (% /!% &' /% &0 %) !) *& -% '%* &0 2!$ &!'9 (! 50 (! *+ !'9 (! 6 $/ %& (! (& .% /* &6 (& '! %0 !$ '!$ !7( .( 7!' 9$ '!% 0! 05 0' $% &$ ,7 (2 !'9 50 !0 $+ (4 5$ /%& 4! '9 (! :* '( &' %$ 7!+ */ !5 0( 0! $& -! $) '%. %'% (0 !, 1! )5 // (& '!$ &!+5 '5 /( !4 (& (/ $' %* &0 ;!v e 9( !$ :: /* $) 9! 9$ 0! ,( (& !(+ %& (!$ 0! $! d() *& 0' /5 )' %& 4! 0' /5 )' 5/ $7 !$ :: /* $) 9e !aw */ \$ !% .#( )72 !i jk jb !0 %& )( ! a% b!% '!0 (: $/ $' (0 !'9 (! () *0 10 '( 6 0! %& '* !0 (. (/ $7 !8 5$ 7%' 1! (7 (6 (& '0 !$ &!%0 '/ %) '0 2!' 9( &! a% %b !%' !) *6 :$ /( 0! '9 (% /!0 '/ 5) '5 /( !%& -% .% -5 $7 71 !$ &2!+ %& $7 71 2!a %%% b!% '!) *6 ,% &( 0! '9 (6 !'* !$ 00 (0 0! '9 (! *. (/ $7 7!) *& -% '%* &0 ;! ! !t ab le 1 .c on tin ue d fr om p re vi ou s pa ge . c om pa ri so n fe at ur es h d (9 2/ 43 /e e c ) b d (2 00 9/ 15 7/ e c ) w fd (2 00 0/ 60 /e c ) m sf d (2 00 8) r ef er en ce s no nco mm er cia l u se on ly e. manea et al.18 weaknesses among directives the comparative analysis of the four directives evidences some weaknesses related to different aspects of their implementation. the absence of precise guidelines for their implementation hampers the possibility of setting up a coherent regional-based ecological network of n2k sites. in addition, management plans related to n2k sites are not mandatory and they are often absent (claudet et al., 2020): this greatly affects the conservation effectiveness, since adequate monitoring activities are not set up, thus leaving the sites often unmanaged and unmonitored, without the possibility to assess the achievement of the conservation objectives they were set for. nature directives may anyway fail in achieving their aims, if they focus only on specific habitats and species, overlooking the overall environmental complexity and ecological connectivity. as for the wfd, the deconstructing and structural approach (see tab. 1; borja et al., 2010) might also weaken the possibility of thoroughly assessing the ecological status of the aquatic environment, by not adequately addressing connectivity aspects. this is something that the msfd tries to avoid by proposing a fully holistic and ecosystem-based approach. however, an effective operationalization of such an approach within all the here considered legislative frameworks is still weak (o’hagan, 2020), despite it would be greatly beneficial to underpin a better protection of the marine environment at the ecosystem level. the msfd and wfd, despite presenting detailed monitoring requirements, do not give indication to member states (ms) on threshold values or baselines to be considered when assessing the level of gens and good ecological status (gecs), respectively. difficulties in establishing current extent and quality of habitats and populations, and in setting acceptable limits to degradation against unknown levels of natural variation, may strongly affect directives’ implementation and lead to incoherency among ms assessment results (dodds et al., 2010). within the diverse monitoring frameworks covered by the four directives there is a noticeable heterogeneity in the terminology used that may create confusion to ms and n2k sites’ managers to define and setting up coherent monitoring strategies. therefore, we highlight the need to harmonize terminology on criteria, quality elements and parameters (see tab. 1). with ecoads, we make a first attempt to define coherent relationships among the considered directives (tables s2-s5). monitoring approaches the hd and bd address the conservation of target species and habitats, thus the parameters they consider for site selection and monitoring are related to the conservation status of these targets, comprising quantitative (range, area, population size) as well as qualitative (structure and functions) criteria plus a forecast for the future (‘future prospects’ parameter) (european commission, 2016a). however, no guidance on how to monitor species and habitats is provided. only guidelines for the reporting of the necessary information for the assessment of sites’ state were delivered (dg environment, 2017). the wfd, differently from the nature directives, provides specific guidelines for monitoring the quality status of water bodies. indeed, it presents very detailed monitoring requirements for both surface and groundwater and lists several steps to carry out for effective monitoring: design of surveillance, operational and investigative monitoring, frequency of monitoring, additional monitoring requirements for protected areas, standards for monitoring of quality elements. the wfd labels the parameters to be adopted for monitoring as quality elements (biological, hydro-morphological, and physico-chemical). this directive asks for a systematic “review of the impact of human activity on the status of surface waters and groundwater”. however, similarly to the nature directives, it does not specifically list the human activities and related pressures to be considered and managed. the msfd articulates the gens in eleven descriptors (table s1) and associated criteria by addressing both species and habitats of priority for conservation. the european commission, in 2017, laid down criteria and methodological standards to give common guidelines to ms to monitor advances towards the achievement of gens through the eu 2017/848. another document was delivered concurrently, the eu 2017/845, to better guide the ms during the second cycle of implementation of their marine strategies. ms should take into account pressures or impacts of human activities in each marine region or subregion, having regard to the indicative lists set out in annex iii. tables s2-s5 report the comparative analysis focalized on the descriptive indicators of the four directives. we use the term descriptive indicators as a common term to refer to the hd and bd parameters, the quality elements of the wfd, and the msfd criteria. since the msfd is the most recent directive and adopts the most holistic approach to monitoring, the comparison of the indicators is developed starting from its criteria. criteria are compared with the descriptive indicators of the other directives to find coherent matches to make a first harmonization attempt. it is noticeable that there is not a complete and specific correspondence between the diverse descriptive indicators, which differ among directives as each jurisdictional instrument addresses specific conservation targets despite their overall shared objectives. indeed, they are aggregated differently: the msfd criteria are articulated upon its descriptors, which identify specific environmental performances (descriptors 1-4) explicitly considering also the human-induced pressures sphere (descriptors 5-11). interno nco mm er cia l u se on ly a transnational marine ecological observatory in the adriatic sea 19 estingly, the msfd is the only directive that reflects, even if indirectly, on species interactions through descriptor 4, related to the food web and the elements from which this is composed. indeed, criteria related to d4 find almost no correspondence with the descriptive indicators of the other directives. the wfd presents a rich set of specific quality elements related to the water bodies’ characteristics, including also inland (e.g., lakes and rivers) and transitional water ecosystems (e.g., lagoons), which are not covered by the msfd. the wfd specifically addresses only some target communities, i.e., macrophytes, phytobenthos, phytoplankton, invertebrates, and fish. it does not include pelagic population beyond phytoplankton and fish components (e.g., cephalopods, reptiles and marine mammals). however, some wfd quality elements find a good correspondence with different msfd criteria, as for example those related to the phytoplankton component and linked to d5. also several wfd chemical and physical elements closely match (e.g., all elements related to the hydrographical conditions) with msfd criteria, even though the terminology and the level of specificity of the indicators are different. regarding the hd and bd, it is possible to observe a stronger correspondence of their parameters with criteria related to msfd descriptor 1 “biodiversity is maintained”. for instance, both nature directives find correspondence with three criteria related to d1: d1c2 population abundance, d1c3 demographic characteristics, and d1c4 distributional range and pattern. the corresponding parameter in hd is “population dynamics of animal and plant species”, while in bd is “trends and variations in population for species”. both nature directives do not consider the set of chemical and physical indicators included in the other two directives (msfd and wfd), being mainly focalized on target species and habitats preservation. although the hd does not look at all species occurring in water bodies (i.e., the aquatic community as a whole), being mainly focused on the conservation status of selected species and habitats, there are some quality elements defined by the wfd and some criteria defined by the msfd that if jointly monitored can be shared and beneficial for n2k sites monitoring, in particular for those monitoring activities that require standardized methodological approaches (e.g., chl-a measurements and nutrient analysis). future perspective for ecoads as transnational meo cross-border cooperation to monitor the marine environment has been highlighted as crucial to support comprehensive monitoring and effective conservation and management (stelzenmüller et al., 2013; azzurro et al., 2014). the need of such an approach has become greater with time since eu member states have been called to integrate the planning and management of their territorial waters into a transboundary approach by year 2020 (european commission 2014). cross-border conservation initiatives have been suggested as powerful tools to foster a balance between conservation and exploitation of natural resources, as well as to reflect the ecological boundaries dictated by marine connectivity (mackelworth, 2012; costello and molina, 2021). in the adriatic sea, where n2k network is under further implementation, several studies have been developed to guide marine conservation initiatives and ecosystem-based management, also in transboundary contexts (gissi et al., 2018; manea et al., 2019; drius et al., 2019; farella et al., 2020 among others). the establishment of the transnational meo ecoads, shared between italy and croatia, represents an opportunity for setting up a common platform where to implement the harmonization and integration of marine environment monitoring to inform both the local and basin-scale conservation strategies that are underway (bax et al., 2018). indeed, in this area an integrated system that coordinates the diverse monitoring efforts is absent (manea et al., 2020). here we focus on the monitoring frameworks entailed within the nature directives and the wfd and msfd carried out in the area. we bring into light many synergies among them as a base for a comprehensive monitoring of the marine environment through space and time to inform the development of ecoads. these synergies are crucial since the monitoring under the nature directives is weakly implemented, often for shortcomings of dedicated funds. we indicate as key the wfd and msfd monitoring programs to investigate n2k sites environmental state, also considering that (as evidenced in section “ecoads”), most of hd and bd target species and habitats are monitored under the msfd. thus, wherever possible, joint monitoring programs should be arranged to include priority conservation sites within their spatial footprints, also to avoid neglecting some marine areas, especially at the interfaces (e.g., land-sea, coastal-offshore waters). this would be particularly relevant to integrate ecological connectivity information in monitoring strategies, even beyond conservation sites to fit a transboundary context (portman and teff-seker, 2017). ecoads is conceived as an instrument able to boost such coordination, thus favouring the integration of these monitoring efforts at multiple scales and the availability of observatory data to inform management and conservation strategies at the adriatic sea region. as a key principle, ecoads addresses the incorporation of ecological connectivity to support the adoption of an ecosystem-based approach to management (jonsson et al., 2021). the legal documentation considered in this study supports this necessity, despite directives barely entail econo nco mm er cia l u se on ly e. manea et al.20 logical connectivity aspects, thus orienting the implementation of this approach only weakly (o’hagan, 2020). for instance, our analysis points out that the trophic webs in the marine environment are extremely poorly characterized by the existing monitoring programs, even though they are an integral part of the connectivity concept (dias et al., 2016), in combination with the state of the species and the whole oceanographic context. to address such weakness, ecoads aims at developing a shared set of monitoring indicators entailing both ecological and oceanographic data, giving relevance to connectivity aspects. indeed, only a monitoring strategy implemented and coordinated at the basin scale, and addressing processes like larvae dispersal and genetic flow, nutrient fluxes, migration patterns and movement in response to environmental variability and climate change and human-induced pressures (toonen et al., 2011; she et al., 2016) can support the incorporation of ecological connectivity in marine monitoring. furthermore, starting from the msfd indications, ecoads will entail the monitoring of anthropogenic pressures’ sources and footprint. indeed, human pressure monitoring is essential to inform management and measure conservation outcomes (dunham et al., 2020). the harmonization of the terminology used to indicate monitored parameters and variables is an additional purpose of ecoads, since the heterogeneous use of terms can severely hamper an effective coordination between monitoring schemes also linked to the fulfilment of the different eu obligations. this effort is grounded on the awareness that the absence of harmonized terminology leads to obstacles in data integration and interpretation, thus inducing redundant research and monitoring efforts and slowing down the scientific cooperation and progress (berners-lee et al., 1999). the adoption of a shared and agreed set of terms to refer to monitoring indicators, linked together in a hierarchical and relational network – within the so called “thesauri” (iso, 2011, 2013) can help overcome the ambiguities associated with data markup and allow digital interpretation and information storage of records (ioc unesco, 2019), thus improving the interoperability and data exchange. as an outcome, in ecoads a harmonized set of monitoring indicators will be distilled starting from those proposed by the discussed directives (tables s1-s5), and it will be integrated in an interactive web application to support the setting up of the monitoring activities needed in selected adriatic n2k sites. the indicators will be further elaborated according to selection criteria, such as their policy relevance, sensitivity to change and feasibility, to evaluate their suitability and priority level for monitoring (schmeller et al., 2018). this set of indicators will also support the ebv and eov frameworks, which have been developed to address the need to track progress towards sdgs and conservation goals (reyers et al., 2017). ecoads, playing the role of collector of different monitoring data sources, can contribute to the ongoing building of long-term datasets to inform the definition of reliable threshold values useful to map and assess the state of the marine environment. moreover, the adoption of the open science approach and the fair principles (european commission 2016b) are addressed to boost transnational data sharing through the implementation of a web portal (manea et al. 2020) acting as a unique access point to the data collected by the observatory. ecoads represents a real opportunity to better operationalize the ecosystem-based approach, since it contributes to implement a holistic approach to marine monitoring and conservation. indeed, within its framework, a core part is taken by the engagement of stakeholders and right holders who might contribute to co-produce knowledge related to the marine environment and the use of its resources. ecoads is not only at the interface between scientific research and environmental management, but also between the scientific world and the civil society. community-based monitoring has been highlighted as a possible effective approach to engage in future monitoring practices (turricchia et al., 2020). running a multi-level participatory process is at the base of the development of ecoads, to design its structure and nourish its conceptual framework. in addition, this process can be of great benefit for the implementation of eu directives, since stakeholders’ contribution is crucial to put into practice environmental policies and to achieve marine conservation goals. acknowledgments all the authors acknowledge the ecoss project (observing system in the adriatic sea: oceanographic observations for biodiversity), an eu 2014–2020 interreg v-a italy-croatia cbc programme id number 10042301. corresponding author: caterina.bergami@ismar.cnr.it key words: marine ecological observatories; transboundary coordination; ecoads; adriatic sea; ecoss project. conflict of interest: the authors declare that they have no known competing financial interests or personal relationships that could have appeared to influence the work reported in this paper. received: 15 april 2021. accepted: 24 june 2021. this work is licensed under a creative commons attribution noncommercial 4.0 license (cc by-nc 4.0). ©copyright: the author(s), 2021 licensee pagepress, italy advances in oceanography and limnology, 2021; 12:9811 doi: 10.4081/aiol.2021.9811 no nco mm er cia l u se on ly a transnational marine ecological observatory in the adriatic sea 21 references acri f, bastianini m, bernardi aubry f, camatti e, boldrin a, et al., 2020. a long-term (1965–2015) ecological marine database from the lter-italy northern adriatic sea site: plankton and oceanographic observations. earth syst. sci. data 12:215-230. aguzzi j, iveša n, gelli m, costa c, gavrilovic a, et al., 2020. ecological video monitoring of marine protected areas by underwater cabled surveillance cameras. mar. pol. 119:104052. allan ij, vrana b, greenwood r, mills ga, roig b, gonzalez c, 2006. a “toolbox” for biological and chemical monitoring requirements for the european union’s water framework directive. talanta 69:302-322. azzurro e, ben souissi sj, boughedir w, castriota l, deidun a, et al. 2014. the sicily strait: a transnational observatory for monitoring the advance of non indigenous species. biol. mar. mediterr. 21:105-106. berners-lee t, fischetti m, 2000. weaving the web: the original design and ultimate destiny of the world wide web by its inventor. harper: san francisco. bastari a, micheli f, ferretti f, pusceddu a, cerrano c, 2016. large marine protected areas (lmpas) in the mediterranean sea: the opportunity of the adriatic sea. mar. pol. 68:165-177. bax nj, appeltans w, brainard r, duffy je, dunstan p, et al., 2018. linking capacity development to goos monitoring networks to achieve sustained ocean observation. front. mar. sci. 5:346. benedetti-cecchi l, crowe t, boehme l, boero f, christensen a, et al., 2018. strengthening europe’s capability in biological ocean observations. future science brief 3 of the european marine board: ostend. available from: https://marineboard.eu/sites/marineboard.eu/files/public/pub lication/emb_fsb3_biological_ocean_observation.pdf biermann f, kanie n, kim re, 2017. global governance by goal-setting: the novel approach of the un sustainable development goals. curr. opin. env. sust. 26-27:26–3. borja a, 2005. the european water framework directive: a challenge for nearshore, coastal and continental shelf research. cont. shelf res. 25:1768-1783. borja á, elliott m, carstensen j, heiskanen as, van de bund w, 2010. marine management–towards an integrated implementation of the european marine strategy framework and the water framework directives. mar. pollut. bull. 60:2175-2186. carr mh, neigel je, estes ja, andelman s, warner rr, largier jl, 2003. comparing marine and terrestrial ecosystems: implications for the design of coastal marine reserves. ecol. appl. 13:90-107. carr mh, woodson cb, cheriton om, malone d, mcmanus ma, raimondi pt, 2011. knowledge through partnerships: integrating marine protected area monitoring and ocean observing systems. front. ecol. environ. 9:342-350. claudet j, loiseau c, sostres m, zupan m, 2020. underprotected marine protected areas in a global biodiversity hotspot. one earth 2:380-384. cossarini g, querin s, solidoro c, 2015. the continental shelf carbon pump in the northern adriatic sea (mediterranean sea): influence of wintertime variability. ecol. model. 314:118-134. costello c, molina r, 2021. transboundary marine protected areas. resour. energy econ. 65;101239. crise a, ribera d’alcalà m, mariani,p, petihakis g, robidart j, et al., 2018. a conceptual framework for developing the next generation of marine observatories (mobs) for science and society. front. mar. sci. 5:318. dg environment, 2017. reporting under article 17 of the habitats directive: explanatory notes and guidelines for the period 2013-2018. available from: http://www.reportingdirettivahabitat.it/public/documents/documenti%20ufficiali%20europei/reporting%20guidelines.pdf dias e, morais p, cotter am, antunes c, hoffman jc, 2016. estuarine consumers utilize marine, estuarine and terrestrial organic matter and provide connectivity among these food webs. mar. ecol. prog. ser. 554:21-34. dodds wk, clements wh, gido k, hilderbrand rh, king rs, 2010. thresholds, breakpoints, and nonlinearity in freshwaters as related to management. j. n. am. benthol. soc. 29:988-997. drius m, bongiorni l, depellegrin d, menegon s, pugnetti a, stifter s, 2019. tackling challenges for mediterranean sustainable coastal tourism: an ecosystem service perspective. sci. total environ. 652:1302-1317. duffy je, amaral-zettler la, fautin dg, paulay g, rynearson ta, et al., 2013. envisioning a marine biodiversity observation network. bioscience 63:350-361. dunham a, dunham js, rubidge e, iacarella jc, metaxas a, 2020. contextualizing ecological performance: rethinking monitoring in marine protected areas. aquat. conserv. 30:2004-2011. european commission, 1992. council directive 92/43/eec of may 21 1992 on the conservation of natural habitats and of wild fauna and flora. official journal of the european communities. european commission, 2000. directive 2000/60/ec of the european parliament and of the council of 23 october 2000 establishing a framework for community action in the field of water policy (water framework directive). official journal of the european union. european commission, 2008. directive 2008/56/ec of the european parliament of the council of 17 june 2008 establishing a framework for community action in the field of marine environmental policy (marine strategy framework directive). official journal of the european union. european commission, 2009. directive 2009/147/ec of the european parliament and of the council of 30 november 2009 on the conservation of wild birds. official journal of the european union. european commission, 2012. faq final 2012-07-27. links between the marine strategy framework directive (msfd 2008/56/ec) and the nature directives (birds directive 2009/147/eec (bd) and habitats directive 92/43/eec (hd)) https://ec.europa.eu/environment/nature/natura2000/marine/d ocs/faq%20final%202012-07-27.pdf european commission, 2014. report from the commission to the european parliament, the council, the european economic and social committee of the regions concerning the governance of macroregional strategies. document 52014dc0284. european commission, 2016a. reporting under article 17 of the habitats directive. report format for the period 2013– no nco mm er cia l u se on ly e. manea et al.22 2018, final version – november 2016. available from: http://cdr.eionet.europa.eu/help/habitats_art17/index_html european commission, 2016b. european commission: open innovation, open science, open to the world – a vision for europe. available from: https://op.europa.eu/en/publicationdetail/-/publication/3213b335-1cbc-11e6-ba9a-01aa75ed71a1/ language-en european commission, 2017. commission directive 2017/845 of 17 may 2017 amending directive 2008/56/ec of the european parliament and of the council as regards the indicative lists of elements to be taken into account for the preparation of marine strategies. document 32017l0845. european commission, 2017. commission decision 2017/848 of 17 may 2017 laying down criteria and methodological standards on good environmental status of marine waters and specifications and standardised methods for monitoring and assessment, and repealing decision 2010/477/eu. document 32017d0848. european commission, 2017. communication from the commission to the european parliament, the council, the european economic and social committee and the committee of the regions an action plan for nature, people and the economy. available from: https://op.europa.eu/en/publication-detail//publication/bb26e7d1-2b26-11e7-9412-01aa75ed71a1 european environment and sustainable development advisory councils (eeac), 2018. working group on fresh water affair 2018. the eu water framework directive. results to date and outlook for the future. available from: https://eeac. eu/wp-content/uploads/2018/09/the-eu-water-frameworkdirective-results-to-date-and-outlook-for-the-future.pdf evans, d, 2012. building the european union’s natura 2000 network. nat. conserv. 1:11. farella g, menegon s, fadini a, depellegrin d, manea e, et al., 2020. incorporating ecosystem services conservation into a scenario-based msp framework: an adriatic case study. ocean. coast. manag. 193:105230. gissi e, mcgowan j, venier c, di carlo d, musco f, et al., 2018. addressing transboundary conservation challenges through marine spatial prioritization. conserv. biol. 32:1107-1117. grodzińska-jurczak m, strzelecka m, kamal s, gutowska j, 2012. effectiveness of nature conservation–a case of natura 2000 sites in poland. in: sladonja b (ed.), protected area management. intechopen. intergovernmental oceanographic commission of unesco, 2019. ocean data standards, vol.4: seadatanet controlled vocabularies for describing marine and oceanographic datasets a joint proposal by seadatanet and odip projects. (ioc manuals and guides, 54, vol. 4.) 31 pp. (ioc/2019/mg/54 vol.4). ostend: iode/unesco. international organization for standardization, 2011. iso 25964-1:2011. information and documentation thesauri and interoperability with other vocabularies part 1: thesauri for information retrieval. geneva: international organization for standardization. international organization for standardization, 2013. norm iso 25964-2:2013. information and documentation thesauri and interoperability with other vocabularies part 2: interoperability with other vocabularies. geneva: international organization for standardization. jonsson pr, hammar l, wåhlström i, pålsson j, hume d, et al., 2021. combining seascape connectivity with cumulative impact assessment in support of ecosystem�based marine spatial planning. j. appl. ecol. 58:576-586. kallis g, butler d, 2001. the eu water framework directive: measures and implications. water pol. 3:125-142. kati v, hovardas t, dieterich m, ibisch pl, mihok b, selva n, 2015. the challenge of implementing the european network of protected areas natura 2000. conserv. biol. 29:260-270. long r, 2011. the marine strategy framework directive: a new european approach to the regulation of the marine environment, marine natural resources and marine ecological services. j. energy nat. resour. law 29:1-44. mack l, attila j, aylagas e, beermann a, borja a, et al., 2020. a synthesis of marine monitoring methods with the potential to enhance the status assessment of the baltic sea. front. mar. sci. 7:823. mackelworth p, 2012. peace parks and transboundary initiatives: implications for marine conservation and spatial planning. conserv. lett. 5:90-98. manea e, di carlo d, depellegrin d, agardy t, gissi e, 2019. multidimensional assessment of supporting ecosystem services for marine spatial planning of the adriatic sea. ecol. indic. 101:821-837. manea e, bongiorni l, bergami c, pugnett, a, 2020. challenges for marine ecological observatories to promote effective gms of natura 2000 network. the case study of ecoads in the adriatic sea, p. 23-39. in: alfaré l, ruoss e (eds.), governing future challenges in protected areas. cnr edizioni. maxwell sm, hazen el, lewison rl, dunn dc, bailey h, et al., 2015. dynamic ocean management: defining and conceptualizing real-time management of the ocean. mar. pol. 58:42-50. mcquatters-gollop a, mitchell i, vina-herbon c, bedford j, addison pfr, et al., 2019. from science to evidence–how biodiversity indicators can be used for effective marine conservation policy and management. front. mar. sci. 6:109. mee ld, jefferson rl, laffoley dda, elliott m, 2008. how good is good? human values and europe’s proposed marine strategy directive. mar. pollut. bull. 56:187-204. micheli f, halpern bs, walbridge s, ciriaco s, ferretti f, et al., 2013. cumulative human impacts on mediterranean and black sea marine ecosystems: assessing current pressures and opportunities. plos one 8:e79889. milliman jd, bonaldo d, carniel s, 2016. flux and fate of riverdischarged sediments to the adriatic sea. adv. oceanogr. limnol. 7:5899. miloslavich p, bax nj, simmons se, klein e, appeltans w, et al., 2018. essential ocean variables for global sustained observations of biodiversity and ecosystem changes. glob. change biol. 24:2416-2433. muelbert jh, nidzieko nj, acosta at, beaulieu se, bernardino af, et al., 2019. ilter–the international long-term ecological research network as a platform for global coastal and ocean observation. front. mar. sci. 6:527. muller-karger fe, miloslavich p, bax nj, simmons s, costello mj, et al., 2018. advancing marine biological observations and data requirements of the complementary essential ocean variables (eovs) and essential biodiversity variables (ebvs) frameworks. front. mar. sci. 5:211. no nco mm er cia l u se on ly a transnational marine ecological observatory in the adriatic sea 23 o’hagan am, 2020. ecosystem-based management (ebm) and ecosystem services in eu law, policy and governance, p. 353-372. in: o’higgins t, lago m and dewitt t (eds.), ecosystem-based management, ecosystem services and aquatic biodiversity. cham: springer. pereira hm, ferrier s, walters m, geller gn, jongman rhg, et al., 2013. essential biodiversity variables. science 339:277–278. portman me, teff-seker y, 2017. factors of success and failure for transboundary environmental cooperation: projects in the gulf of aqaba. j. environ. policy plan 19:810-826. rayner r, jolly c, gouldman c, 2019. ocean observing and the blue economy. font. mar. sci. 6:330. ravaioli m, bergami c, riminucci f, langone l, cardin v, et al., 2016. the ritmare italian fixed-point observatory network (ifon) for marine environmental monitoring: a case study. j. oper. oceanogr. 9:202–214. reyers b, stafford-smith m, erb kh, scholes rj, selomane o, 2017. essential variables help to focus sustainable development goals monitoring. curr. opin. environ. sustain. 26:97105. rouillard j, lago m, abhold k, röschel l, kafyeke t, et al., 2018. protecting aquatic biodiversity in europe: how much do eu environmental policies support ecosystem-based management? ambio 47:15-24. scarcella g, grat, f, raicevich s, russo t, gramolini r, et al., 2014. common sole in the northern and central adriatic sea: spatial management scenarios to rebuild the stock. j. sea res. 89:12-22. schmeller ds, weatherdon lv, loyau a, bondeau a, brotons l, et al., 2018. a suite of essential biodiversity variables for detecting critical biodiversity change. biol. rev. 93:55-71. she j, allen i, buch e, crise a, johannessen a, et al., 2016. developing european operational oceanography for blue growth, climate change adaptation and mitigation, and ecosystem-based management. ocean sci 12:953–976. šepić j, vilibic i, 2011. the development and implementation of a real-time meteotsunami warning network for the adriatic sea. nat. hazards earth syst. sci. 11:83-91. šepić j, vilibic i, denamiel c, mihanović h, muslim s, et al., 2017. towards understanding and operational early warning of the adriatic meteotsunamis: project messi. proceedings 1st workshop on waves, storm surges and coastal hazards. stelzenmüller v, breen p, stamford t, thomsen f, badalamenti f, et al., 2013. monitoring and evaluation of spatially managed areas: a generic framework for implementation of ecosystem based marine management and its application. mar. pol. 37:149-164. toonen rj, andrews kr, baums ib, bird ce, concepcion gt, et al., 2011. defining boundaries for ecosystem-based management: a multispecies case study of marine connectivity across the hawaiian archipelago. j. mar. biol. 2011:460173. united nations environment programme (unep), 2011. taking steps toward marine and coastal ecosystem-based management an introductory guide. available from: https://www. unep.org/resources/report/taking-steps-toward-marine-andcoastal-ecosystem-based-management-introductory united nations environment programme (unep), 2019. frontiers 2018/19: emerging issues of environmental concern. available from: https://www.unep.org/resources/frontiers201819-emerging-issues-environmental-concern no nco mm er cia l u se on ly layout 1 introduction the term diadromy describes migrations between freshwater and marine environments (myers, 1949; mcdowall, 1988; mcdowall, 1992). diadromous species include less than 3% of the world fish fauna (eschmeyer and fong, 2016), among which several ones are economically and culturally important, such as freshwater anguillid eels and salmons (chapman et al., 2012). catadromous fish are characterized by a complex life cycle where fish breed in the ocean and growth in continental coastal and/or inland waters (mccleave, 2001), as seen in anguillids (tesch, 2003; elliot et al., 2007). main ecological services provided by catadromous fish consist for example in the provision of food, and in the regulation of ecosystem functions by transporting nutrients and linking different biomes (druineau et al., 2018a). globally, these animals have been appreciated for human consumption showing a relevant economic interest (costa-dias et al., 2009; feunteun and laffaille, 2011). catadromous fish can be also used as indicators of environmental quality and functionality (smith et al., 2016). for instance, they are also commonly used as a metric in the assessment of water bodies ecological status in the european water framework directive (delpech et al., 2010) or as bio-indicators of water quality (amara et al., 2009), reflecting both habitat longitudinal connectivity and habitat quality. in this context, an exiguous number of catadromous fish species are identified as ‘umbrella species’ in order to ensure the protection of these species and their habitats (rochard et al., 2009). they are also magnified by many cultures, foster a belonging sense, and support million-dollar fishing (garman, 1992; close et al., 2002; montgomery, 2003; chasco et al. 2017; noaa, 2017). because of this general interest, catadromous fish are object of studies in all their dimensions (drouineau et al., 2018b) and strongly linked to research questions associated with animal migration (secor, 2015; morais and daverat, 2016). catadromous fish use along their migration pathways a variety of habitats and face many diverse environmental threats (mcintyre et al., 2016). in europe, as observed for most migratory animals (sanderson et al., 2006; wilcove and wikelski, 2008), a worldwide decline of migratory fish has been recorded at least since the beginning of the 20th century (béguer et al., 2007; wolter, 2015; lambert et al., 2019). the causes are numerous and likely cumulative (e.g., obstacles to migration, deterioration in essential habitat and water quality, unsustainable fisheries, parasite introductions), although quantitative evidence has been rarely demonstrated (dekker and casselman, 2014). as a result, many catadromous species are now classified as rare, endangered, or extinct, in the iucn red list (iucn, 2019). there is, therefore, an urgent need to develop approaches that provide reliable quantification of the specific impacts of the different anthropogenic pressures acting on catadromous species. this would help support the implementation of effective mitigation measures and provide adequate tools for national and international regulation around the world. among the most iconic catadromous species, the european eel anguilla anguilla (linnaeus, 1758) has been the focus of many studies (e.g., dekker, 2003a; bonhommeau et al., 2008; kettle et al., 2011; baltazar-soares et al., 2014; schiavina et al., 2015; aalto et al., 2016; righton et al., 2016; bornarel et al., 2018; bevacqua et al., 2019; dekker, 2019). the life cycle of the european eel a. anguilla has stimulated great curiosity and interest since at least the 4th century bc, where already some important greek philosophers like aristotle hypothesized on the origin of this species, which remained enshrouded in mystery for millennia. the recent interest increase in eel biology is primarily linked to conservation issues. therefore, to implement our knowledge about the main natural and anthropogenic threats to its survivorship and identify possible solutions to preserve it, there is an urgent need to gain further insights into a. anguilla life-history. review hard times for catadromous fish: the case of the european eel anguilla anguilla (l. 1758) cinzia podda,*1 francesco palmas,1 antonio pusceddu,1 andrea sabatini1 1department of life and environmental sciences, university of cagliari, via fiorelli 1, 09126 cagliari, italy abstract catadromous fish species can be defined important organisms for their ecological, economical, and cultural value. because of a complex life cycle, catadromous fish species are exhibited to the cumulative effect of multiple anthropogenic threats that resulted in worldwide decline since the beginning of the 20th century. among the most iconic catadromous species, the european eel anguilla anguilla has aroused considerable interest, and to date, many aspects of its life cycle remain relatively unknown. although conspicuous efforts by the research to ensure the perpetuation of the species were conducted, the identification of the best tools to reduce the threats that affect eels remains challenging. in this narrative review, the state of the knowledge and main threats about the life cycle, the habitat occupancy, the recruitment, and migration patterns of the european eel have been reported. no nco mm er cia l u se on ly c. podda et al.48 in this narrative review, we present the state of the knowledge about the life cycle, habitat occupancy, recruitment, and migration patterns of the european eel and about the major threats most likely have contributed to the decline of eels. life cycle, habitat occupancy and migration patterns of the life cycle of the genus anguilla: common features eels of the genus anguilla (schrank, 1798) are the only genus of anguilliformes with a catadromous life cycle (mcdowall, 1988). the life cycle of anguillid eels involves five developmental stages: leptocephalus (larvae), glass eel (transparent juvenile stage), elver (pigmented juvenile stage), yellow eel (immature adult) and silver eel (partially mature adult) (bertin, 1956; tesch, 1977; cresci, 2020). the larval stage duration varies in different regions worldwide and can last from several months to some years, according to the species and biogeographic region (tsukamoto, 1990; lecomte-finiger, 1992; cheng and tzeng, 1996; arai et al., 1999; 2001; 2003; wang and tzeng, 2000; marui et al., 2001; robinet et al., 2003; robinet et al., 2008; reveillac et al., 2008; 2009; bonhommeau et al., 2010; han et al., 2016; 2019; hewavitharane et al., 2020). the larval phase suffers high mortality thereby influencing recruitment success (cushing, 1990; durant et al., 2007). after metamorphosis into glass eels, juveniles leave oceanic waters, starting the upstream migration crossing coastal waters (tesch, 2003; cresci, 2020). glass eels represent the recruitment phase to continental waters (ices, 2011) and constitute the natural source of supply of the species because its artificial reproduction is not yet possible (pedersen and ramussen, 2016). glass eels develop into elvers and settle as yellow eels for many years (about 525 years) in coastal and inland water habitats (e.g., estuaries, rivers, streams, ponds, and lakes) (tesch, 2003; cresci, 2020). after this trophic phase, eels start the downstream migration during the silver eel stage (tesch, 2003) that is initiated by endogenous and exogenous signals that coincide with optimal conditions for successful migration (sandlund et al., 2017). the migration peaks in rivers properly occur during rainfall events associated with flow pulses, affecting water velocity, turbidity, and conductivity (cullen and mccarthy, 2003; durif et al., 2008; drouineau et al., 2017). once gonad maturation starts, eels run downstream mainly at night, during rising river flow phases (behrmann-godel and eckmann, 2003), which also provide protection against predation and reduce the swimming energy cost to return to the offshore spawning area (tesch, 2003; sandlund et al., 2017; cresci, 2020). habitat occupancy the european eel a. anguilla is a panmictic species (palm et al., 2009; enbody et al., 2021) distributed across most of the coastal countries in europe and north africa and spanning the entire mediterranean basin (ices, 2018). because the complex life cycle, the cryptic behavior, and body shape features of eels, results hard to find appropriate and standardized sampling technique for the monitoring of the european eel in several aquatic environments (naismith and knight, 1990; lasne and laffaille, 2009). furthermore, many aspects of the resident stage of eels in freshwaters during their growth phase are still insufficiently understood such as ecology in terms of space and time use (feunteun et al., 2003; imbert et al., 2010). the habitats occupancy can be investigated through the otolith microchemistry used to determine the type of habitat of individuals throughout their life, primarily using the strontium (sr) to calcium (ca) ratio to distinguish freshwater phases from brackish and seawater phases (tsukamoto and aoyama, 1998; arai et al., 2006; shiao et al., 2006; lin et al., 2011 arai et al., 2019). more recently, other elements, such as barium (ba) and manganese (mn), have been used to assess finer-scale movement patterns (benchetrit et al., 2017). this technique constitutes a reliable tool for the assessment of habitat use and growth throughout the entire life span between freshwater and saline waters (clément et al., 2014). experimental electrofishing has been recognized an efficient sampling method to catch eel in freshwaters despite some limits (e.g., deep waters) (laffaille and rigaud, 2008), while fishery-based time-series are usually utilized to assess eels’ temporal trends (ices, 2020). glass eel fisheries are carried out in the estuaries, or under dams, to study the natural abundance of glass eels in time and space (dekker et al., 2003b). several dipnet types are applied, on foot or using boats (aubrun, 1986), trawls (aubrun, 1987), stow nets (weber, 1986), and fyke nets (ciccotti et al., 2000). fisheries for yellow and silver eels foresee a wide range of gears that include nets, spears, pots, hooks, in coastal areas, lagoons, rivers, lakes, and streams (dekker et al., 2003b). several studies used telemetry to investigate individual movement patterns, site fidelity, habitat use and home range exploitation in relation to seasonal and environmental factors (e.g., ovidio et al., 2013; barry et al., 2015; piper et al., 2017; trancart et al., 2018; dorow et al., 2019; piper et al., 2019; teichert et al., 2020). furthermore, diel, and seasonal phenology and the effect of environmental drivers on non-migrant eel movements were investigated using acoustic camera to evaluate the presence of eels swimming toward the inland waters (lagarde et al., 2021). studies on the presence of eels’ population were conducted also with visual observation no nco mm er cia l u se on ly hard times for the catadromous european eel 49 in inland waters (e.g., lakes and reservoirs) (rossier, 1997; schulze et al., 2004). another tool that could support to understand the eels’ habitat occupancy is represented by the environmental dna (edna) analysis (knudsen et al., 2019). edna assays for target species and edna metabarcoding are both promising techniques for establishing species presence from environmental samples (taberlet et al., 2012; evans et al., 2016; deiner et al., 2017). these indirect methods are cheap to implement at a large scale and can be used to quickly establish the spatial distribution of a target species (atkinson et al., 2018; bracken et al., 2019). instead, when it is difficult to assess the presence of a species because the species couldn’t simply be present, direct methods (fish tagging) or physical survey assessments may be more appropriate (kemp and o’hanley, 2010). juveniles’ migration and orientation migration mechanisms, including orientation, behavior and route architecture throughout the entire life of anguillid eels have been revealed by means of the recent advanced technologies like agent-based model, abm, particle tracking model of upstream migrating juvenile eels (padgett et al., 2020; benson et al., 2021), motion analysis of glass eels (eldrogi et al., 2018), tiny acoustic transmitters (fischer et al., 2019; mueller et al., 2019; liss et al., 2021), satellite tracking for migrant adults (e.g., aarestrup et al., 2009; westerberg et al., 2014; wysujack et al., 2015; amilhat et al., 2016; righton et al., 2016 for the european eel; manabe et al., 2011; higuchi et al., 2018 for the japanese eel anguilla japonica; schabetsberger et al., 2013; 2015; 2019 for pacific eels a. marmorata and a. megastoma; beguer-pon et al., 2015 for the american eel anguilla rostrata). to reduce migration energy costs (forward and tankersley, 2001; bureau du colombier et al., 2007; edeline, 2007), juveniles catadromous species are transported in continental waters by entering the water column during floodtides and descending to the bottom during ebbtides using flood tide transport (ftt) to migrate through estuaries and thus colonize catchments (forward and tankersley, 2001). but in absence of this condition, an alternative migratory tactic to undertake upstream migration reckon on an active swimming running after salinity gradient (cresci, 2020), and using earthy and green odors as attractant (sola and tongiorgi, 1996). for instance, several authors showed that chemical cues (e.g., green odors, amino acids, and bile salts) such as freshwater plumes and salinity gradients transporting inland odors into estuaries can guide estuarine juveniles’ migration (tosi et al., 1988; tosi et al., 1989; crnjar et al., 1992; tosi and sola, 1993; sola, 1995; atta et al., 2013). indeed, eels possess one of the most sensitive olfactory systems among fish, and olfaction plays a central role in their life (huertas et al., 2008). glass eels, for example, are attracted by inland odors, derived from the decomposition of detritus associated with the flora and microfauna in freshwater (sorensen, 1986). among inland odors, geosmin (trans1,10-dimethyltrans9-decalol) play a role in attracting glass eels (tosi and sola, 1993; sola, 1995). in addition, it would seem that geosmin operates as an attractant in freshwater and as a repellent in sea water (tosi and sola, 1993). moreover, glass eels’ migratory behavior may be also affected by physiological changes, alterations of locomotor activity, and decreasing of body condition (edeline et al., 2007). social interactions represent a selective element for the migration and timing of glass eels’ settlement linked to specific habitat survival and growth (edeline et al., 2009). some studies provided evidence that thyroid hormones are involved in glass eel migration (edeline et al., 2004; 2005). decreasing levels of thyroid hormones in juvenile eels might explain the decreasing rate of development and the decreasing propensity to migrate during the transition from the leptocephalus larval to the elver stage (jegstrup and rosenkilde 2003). this hypothesis is corroborated in subadults of the american eel a. rostrata, in which elevated t4 plasma levels are correlated with increased locomotion activity (castonguay et al., 1990). furthermore, european glass eels’ river colonizers exhibit increased thyroid hormone concentrations when compared to estuarine migrants (castonguay et al. 1990; edeline et al., 2004) suggesting a switch from a high migrating to settling behavior strongly linked to thyroid hormones production. the migration of catadromous fish can also be explained with the ‘pheromone hypothesis’, according to which fish can release particular odors into the water (pheromones, likely amino acids; crnjar et al., 1992) functioning as attractants for conspecifics (schmucker et al., 2016). the attractive power of these cues is stage dependent in the eels, it is more accentuated on glass eels and gradually disappears in pigmented glass eels and elvers likely due to physiological and behavioral changes, alternative cues possibly become main attractants (schmucker et al., 2016; galbraith et al., 2017). mechanisms regulating glass eels’ orientation are most likely innate and involve geomagnetic-based compass mechanisms based on the inclination and intensity of the magnetic field (naisbett-jones et al., 2017). however, leptocephali stage present in the sargasso sea may not possess that same magnetic sensing ability as the glass eels because substantial body rearrangements and related physiological changes (tesch, 2003; baltazar-soares and eizaguirre, 2017). more recent studies confirmed that glass eels can orient their migration using the earth’s magnetic field (cresci et al., 2017, cresci et al., 2019a) and lunar cues (cresci et no nco mm er cia l u se on ly c. podda et al.50 al., 2019b), as a reference to imprint a memory of tidal currents in estuaries and to facilitate position holding and upstream migration (cresci et al., 2019b). however, although many individual pieces of the complex puzzle of glass eels’ orientation and migratory behavior have been elucidated, a holistic mechanism to discriminate how they migrate from the continental slope to estuaries and whether this path is memorized until returning to the sea in the adult stages is still far from being identified. adult migration behavior the spawning migration patterns of the european eel in the atlantic ocean have been studied due to their long distances (about 2000–8000 km) (schmidt, 1922; miller et al., 2019). the long migration paths are notable because the amplitude of their scale and the excellent ability to trace the birth location using an unrevealed combination of sensory cues (mccleave and kleckner, 1985). mark–recapture studies have been used to evaluate eels’ home range, habitat preferences, diel and seasonal movements (jellyman et al. 1996; oliveira, 1997; laffaille et al., 2003). direct observations of the migratory behavior of yellow and silver eels were made using electronic tags (e.g., amilhat et al., 2016; righton et al., 2016; béguer-pon et al., 2018; dorow et al., 2019; teichert et al., 2020; trancart et al., 2020). telemetry represents a reliable method to study the spatial ecology of eels, providing the opportunity to track fish in real time or from archived data to remote receivers, enabling data to be retrieved without recapturing the tag (torstad et al., 2013). the development and miniaturization of pop-up satellite archival tags have allowed the tracking of silver eels at sea, unravelling part of the mystery surrounding the oceanic migration of anguillid eels (jellyman and tsukamoto 2002; aarestrup et al. 2009; béguer-pon et al. 2015, amilhat et al., 2016). fundamental telemetry studies on silver eels investigated their migration from freshwaters to oceanic spawning areas, including survival, progression rate and behavioral and external physical factors associated with migration. to date, no telemetry studies on eels have been used with other available physiologically oriented sensors, such as electrocardiography or electromyography (cooke et al. 2013), or any other environmental sensors, such as conductivity or oxygen, while tracking eels in the wild. recently, some studies have shed light on the possible effects of global change in eels’ migration patterns: climate change and warming related thermal and hydrological modifications of aquatic ecosystems could delay or bring forward silver eels (verreault et al.,2012) and alter patterns of glass eels’ migration (moore and jarvis, 2008). migratory phenology and habitat change as affected by current climate change should therefore be a priority of future studies. eels’ recruitment dynamics success and extent of eels’ recruitment depends both on global (knights, 2003; kettle and haines, 2006; bonhommeau et al., 2008; pacariz et al., 2014; gutierrezestrada and pulido-calvo, 2015; bornarel et al., 2018) and local factors, whose interaction modulate spatial and temporal dynamics of recruits entering brackish environments and freshwaters (gascuel et al.,1995; arribas et al., 2012; harrison et al., 2014; trancart et al., 2014; aranburu et al., 2015). recruitment dynamics at the local scale can vary daily, seasonally, and annually (bru et al., 2009; laffaille et al., 2007; zompola et al., 2008; arribas et al., 2012, podda et al., 2020), are well known for atlantic estuaries and rivers of europe (beaulaton and castelnaud, 2005; harrison et al., 2014), and relatively less known for the estuaries located in the southernmost distribution area (arribas et al., 2012). although only one cohort recruits each year (desaunay and guerault, 1997), glass eels arrive in different waves from different routes (boëtius and harding, 1985). the preference of glass eels for freshwater or brackish water varies with the body condition and the timing of arrival to the continental shelf (edeline et al., 2005). reconstructions of exact hatching site and migration routes of the larvae and glass eel, based on mere analyses of recruitment and sampling data (dekker, 1998; lecomte-finger, 1992), have been carried out since the early 20th century (boëtius and harding, 1985; van ginneken and maes, 2005; westerberg et al., 2018). most of the available multi-year temporal series on glass eels’ recruitment to european estuaries is based on fishery and/or scientific surveys, however pluriannual fishery independent studies are very scarce. fishery databased glass eels’ recruitment estimates generally suffer from sampling (methods and protocols) and temporal biases (fishing season). for example, in europe, most surveys to estimate recruitment rates have been conducted in rivers or estuaries, where the eels’ dispersion is influenced by the riverbed or river mouth width, allowing easier glass eel samplings (adam et al., 2008; bru et al., 2009; zompola et al., 2008). models like the glass eel recruitment estimation model one (gerem)(drouineau et al. 2016) estimated the annual glass eel recruitment at different spatial scales, providing a recruitment index to robustly compare spatial variation trends, with large biases for specific regions where data are scarce or not existent (e.g., north africa, eastern mediterranean, and the baltic sea). moreover, it must be noticed that an accurate knowledge of the physical-chemical characteristics of the surveyed environments is also needed to properly assess movement and distribution of eels in both the biomes hosting their life cycle (adam et al., 2008). this need creno nco mm er cia l u se on ly hard times for the catadromous european eel 51 ates a significant challenge: precise information on eels’ numbers entering inland waters and moving through the biomes would be collected to understand recruitment dynamics, but, yet it is hard to be obtained because of the complex, often unpredictable, environmental variability of shallow water ecosystems that can mask natural patterns at the relevant spatial scales. implementing standardized data collection programmes of glass eels’ abundance should be therefore a major investment of future research and stock assessment protocols. threats to a. anguilla the global status of the eel is primarily a consequence of a prolonged decline of its recruitment across the entire distribution area (moriarty and dekker, 1997; ices, 2020 and author therein). many factors have been identified as recruitment shortor medium-term drivers but, so far, it has been difficult to reach clear conclusions about what are the primary drivers of its decline. multiple environmental factors (e.g., river flow, changes in the north atlantic oscillation, warming of sea surface temperature, currents) probably affected the documented decline (e.g., gandolfi et al., 1984; domingos, 1992; elie and rochard, 1994; de casamajor et al., 1999; prouzet, 2002; jellyman and lambert, 2003; knights, 2003; polyakov et al., 2005; bouvet et al., 2006; bureau du colombier et al., 2007; friedland et al., 2007; laffaille et al., 2007; adam et al., 2008; bonhommeau et al., 2008; crivelli et al., 2008; kettle et al., 2008; zompola et al., 2008; miller et al., 2009; durif et al., 2011; kettle et al., 2011; arribas et al., 2012; baltazar-soares et al., 2014; hanel et al., 2014; milardi et al., 2018; podda et al., 2020). while eels are still a common species throughout europe, their stocks have been declining rapidly during the last 40-50 years (dekker, 2016). the decline of the eel global stock affects indeed its entire geographical range, also concerning the southern part of its distribution area, as documented by a concurrent decline in glass eels’ recruitment, as well as by contracting local stocks in the mediterranean region (ciccotti, 2005; aalto et al., 2016; amilhat et al., 2016). silver eels’ abundance decreased by as much as 90% between 1975 and 2010 (bevacqua et al., 2015) with human mediated activities being a contributing factor to this decline (calles et al., 2010; feunteun, 2002; piper et al., 2013). it is known that a combination of natural causes and anthropogenic pressures has been impacting both the eel stock and its habitats (jacoby et al., 2015; miller et al., 2016; drouineau et al., 2018b). the european eel is subjected to fishing activities at all continental life stages (from juveniles to adults) and high fishing mortality estimated over the entire life cycle suggests that overfishing represents one of the main threats for the survival of the entire eel population (fao, 2007). furthermore, all commercial production of a. anguilla (intensive and extensive farming, commercial and recreational fishing) depends on the exploitation of wild stocks (juveniles to supply farms, adults for fishing) (ospar, 2010). to deal with this problem there are various regional management measures currently undertaken to regulate european eel fisheries. principal conservation measures in place for glass, yellow and silver eels include a ban on commercial fishing of glass eels, gear regulations, quotas, closed seasons, licenses for fishing, size limits, free gaps in weirs and requirements for elver passes (ringuet et al., 2002). other pressures play an important role in the decline of the european eel, and include also habitat loss, water pollution, parasitism, and migration obstacles (dams, weirs, pumping stations) (e.g., baltazar-soares et al., 2014; culurgioni et al., 2014; 2015; bevacqua et al., 2015; aalto et al., 2016; dekker and beaulaton, 2016). these factors affect european eels most in the continental phase of their life cycle, while environmental factors, such as climate change, mostly influence their oceanic phase (drouineau et al., 2018b). however, as eels can spend most of their life in freshwater (tesch, 2003) the environmental stressors affecting their life in this biome needs to be studied thoroughly. in 2007, the european commission developed a specific legislation (council regulation (ec) no. 1100/2007) to protect eels (european commission, 2007). european eel has been listed also in appendix ii of the convention on international trade in endangered species (cites, 2020) and in appendix ii of convention for the conservation of migratory species (cms) (cites, 2020). most recently, the international union for conservation of nature (iucn) has recently classified the european eel as critically endangered (iucn, 2014; pike et al., 2020). the stock of the european eel is currently at its historical minimum. for more than half a century, stock abundance and fishing yield have declined by about 5% per year, to less than 10% of its historical level (dekker, 2003a; 2004; ices, 2019). from 1980 to 2010 recruitment of young eel (glass eel) from the ocean towards the continent dropped consistently by approximately 15% per year, to 1%-10% of its former levels (dekker, 2000; ices, 2020). the causes of these downward trends are not clear, and, consequently, efficient remedies and mitigation measures are hard to design (dekker, 2016). hence, the dynamics of the population are only marginally known (dekker, 2004) in the current relatively well-documented years, and even more so for the decades during which the stock declined (dekker, 2016). to fill these gaps of knowledge, monitoring programmes have been established, and models of stock dynamics also developed (de leo et al., 2009; walker et al., 2013). many discoveries were made in the 20th century about no nco mm er cia l u se on ly c. podda et al.52 the behavior and movement ecology of this species. extensive sampling programs were conducted in the atlantic ocean to understand the horizontal and vertical movement of eel leptocephali (hanel et al., 2014; miller et al., 2015), and direct observations of the migratory behavior of yellow and silver eels were made using telemetry (amilhat et al., 2016; righton et al., 2016; béguer-pon et al., 2018). however, less is known about the migratory behavior of glass eels during their complex journey from the continental slope to estuaries. general threats to the survivorship of a. anguilla span across their entire home range including either freshwater, marine coastal, and oceanic habitats. threats to reproducing stocks in freshwater are, obviously, conceivably more of concern. freshwater ecosystems are threatened habitats by multiple human disturbances (vörösmarty et al., 2010), which are expected to affect future species ranges (comte et al., 2016; radinger et al., 2016). from a legislative perspective, the presence of obstacles to river flow is important for determining the hydromorphological status of a river in terms of hydrological regime, continuity, morphological condition, and ecological flows (eu water framework directive (wfd) 2000/60/ec; moccia et al., 2020). indeed, in recent years, there has been a growing interest about ecological consequences of river fragmentation by physical obstacles (garcia de leániz, 2008; januchowski-hartley et al., 2013; kroon and phillips, 2016; birnie-gauvin et al., 2017; jones et al., 2019). recent estimates of fragmentation suggest that 63% of rivers worldwide are no longer free-flowing and that half of all rivers reaches have diminished connectivity (grill et al., 2019). rivers’ flows have been regulated for the purposes of flood protection, navigation, and agricultural development, as well as for electricity production and other human uses. however, these regulations have overall degraded river geomorphological and hydrological conditions (e.g., by the fragmentation of river networks and generating a deficit of sediment transport) (alexander et al., 2012; grill et al., 2019). in river ecosystems, fragmentation due to dam building and changes to river flows due to drought may affect river continuity and can be considered a key driver of the anthropocene biodiversity crisis (meybeck, 2003; dudgeon et al., 2006; zarfl et al., 2015). river connectivity interruption threatens ecosystems’ structure and functioning by hindering movements of migratory species, the exchange of individuals and of genetic information between populations (wofford et al., 2005; raeymaekers et al., 2008;), altering aquatic habitats, flow, and sediment transport regimes (bunn and arthington, 2002). disruption of natural movements can affect the extent, viability, and persistence of native aquatic species, and has caused a decline in the distribution and abundance of many fish populations, including eels (feunteun, 2002; burkhead, 2012; katz et al., 2013). in this regard, we stress here that catadromous fish are declining worldwide, also because of direct and indirect effects generated by dams building (shields et al., 2005; collas et al., 2018. in particular dams mediated river fragmentation limits fish dispersal and likely increases their extinction risk (carvajal-quintero et al., 2017; dias et al., 2017). for example, hydroelectric dams can cause injury, direct mortality, delays in migration times, and inhibit downstream migration in a. anguilla (behrmann-godel and eckmann, 2003; durif et al., 2003; winter et al., 2006; bruijs and durif, 2009). downstream passage at non-powered dams (i.e., dams not equipped with turbines) can have minor impacts, as the passage of fish through them is usually safe (besson et al., 2016), but anyway can delay migration (larinier, 2000; larinier and travade, 2002; besson et al., 2016) and result in lower (20%) annual migration rates when compared to equivalent non-obstructed rivers (feunteun et al., 2000; acou, 2006). a high level of connectivity between habitats in a river system and between a river and the sea is vital for sustaining healthy stream fish populations and assemblages that migrate among several habitats, for suitable feeding, spawning, and refuge conditions (lucas and baras, 2001; maitland, 2003; carlsson et al., 2004; perkin and gido, 2012; van puijenbroek et al., 2019): this holds conceivably true particularly for the survivorship of catadromous fish that migrate across different aquatic biomes. eels can climb along waterfalls and weirs of hydropowers (byrne and beckett, 2012). nevertheless, most counteracting methods applied to mitigate negative effects of dams on fish migration, do not grant success for all migratory fish upstream, and even if they do, successful catadromous fish species can encounter unfavorable habitat conditions in reached reservoirs (larinier, 2001; van puijenbroek et al., 2019). upstream migration in presence of dams may be also delayed given the required time to obtain further fish passages (larinier, 2001; lucas and baras, 2001; brink et al., 2018). moreover, because general upstream effects of dams increase with the size of the dam and reservoir (birnie-gauvin et al., 2017; brink et al., 2018), large dams, usually, tend to be more harmful than smaller barriers. downstream migration in presence of dams can enhance mortality due to predation in reservoirs and passage in hydropower turbines or spillway (larinier, 2001; wilkes et al., 2018). hence, independently of the movement direction and of the presence of fish transposition devices, dams can severely impair catadromous fish movement and, thus, are partly responsible for the decline of catadromous species and, in particular, of eels (calles et al., 2010; feunteun, 2002; piper et al., 2013). widespread eel ladders could aid upstream migration, although to date, few efficiency assessments of their efficiency exist (jellyman and arai, 2016). alterno nco mm er cia l u se on ly hard times for the catadromous european eel 53 native approaches to the capture and the transfer of adult eels downstream of a barrier are also used worldwide (ices, 2016; jellyman and unwin, 2017; béguer-pon et al., 2018). a management of the water regimes alterations of the dams during the fish migratory peaks could be also an effective measure (boubee et al., 2001; trancart et al., 2013), but they can be complicated if they are not predicted to limit the economic loss (teichert et al., 2020). success and limits of eels restocking practices restocking practices are used for conservation, protection, or recovery purposes of endangered species and to increase the productivity of fish stocks (fao, 2003). among the actions undertaken to recovery the european eel population, restocking practices in continental systems where natural recruitment is low or absent are still underdeveloped (moriarty and mccarthy, 1982; andersson et al., 1991; wickström et al., 1996; pedersen, 1998; simon and dörner, 2014; ovidio et al., 2015; hanel et al., 2019). eels’ restocking initiated in europe before to the 20th century and has been done for decades across the entire continent (wickström et al., 1996; moriarty and dekker, 1997; psuty and draganik, 2008; dekker and beaulaton, 2016). the release of glass eels in closed catchments can efficiently support local eels’ production and as well as promote local employment (wickström et al., 1996; pedersen, 2000; rosell et al., 2005; psuty and draganik, 2008). moreover, among the conservation measures conceived for inland waters that are distant from the sea, restocking is the only solution that enhances the local stocks simon et al., 2013; ovidio et al., 2015; matondo et al., 2019). considering that a proportion of stocked eels needs to escape as silver eels, contrasting for example hydropower-induced mortality during the downstream migration (winter et al., 2006), restocking is probably the best long-term plan to meet the silver eels’ escapement target in the eel recovery plan of the european union. to date, the success of eels’ artificial reproduction in captivity has not yet been totally obtained, therefore, domestication and aquaculture may represent an effective tool to satisfy purchaser requests and to preserve natural stocks (guarniero et al., 2020). however, this species represents a true challenge for breeding and production (e.g., egg quality, fertilization rate, and larval survival are the main challenges). wild-caught glass eels and elvers represent the only supply of restocking, that can be translocated from estuaries to rivers with low or without natural immigration (pedersen et al., 2000; matondo et al., 2019). in their new freshwater environments, restocked young eels can survive, grow, and mature into silver eels that, ultimately, display a seaward migration behavior that is similar to the one exhibited by naturally recruited wild eels (e.g., shiao et al., 2006; ovidio et al., 2015; kullmann and thiel, 2018; matondo et al., 2019; felix et al., 2020). nevertheless, it is yet to be demonstrated whether restocking is an efficient measure to restore the eels’ stocks and also to produce new mature individuals that could successfully contribute to the successive spawning stocks (westin, 1998; 2003; prigge et al., 2013; westerberg et al., 2014). moreover, further studies are also needed to assess the impact of restocking practices on the future sexual differentiation of the restocked individuals (geffroy and bardonnet, 2015; ovidio et al., 2015). restocked eels’ long-term survival is also still debated (westin, 1998; 2003; prigge et al., 2013; westerberg et al., 2014). the size and stage of restocking material (glass eel vs. yellow eel), their origin (cultured vs. wild eels), their health status (e.g., parasites, infections, deseases), and the trophic status of the water body may altogether influence the restocking yield (prigge et al., 2013; pedersen et al., 2016; ovidio et al., 2015). the annual growth in length and the survival rates of restocked eels vary strongly among different recipient environments, and depend upon the characteristics of rearing location, the wild origin of reared eels, and the stage used (juvenile vs. adult eels) (bisgaard and pedersen, 1991; pedersen, 1998; lin et al., 2007; simon et al., 2013; simon and dörner, 2014). recent studies reported that natural mortality of restocked populations decreases with increasing individual body mass and, thus, restocking carried out with larger eels resulted in a better survival rate and, consequently, in a higher yield (pedersen et al., 2016). more recent experiments showed that restocked eels have an initial delay of their downstream migration, and those recaptured eels have lower body length and weight, likely attributable to their allochthonous origin (prigge et al., 2013). interestingly, however, both restocked and farmed eels show similar migratory behaviors and routes during spawning migrations in the open ocean (westerberg et al., 2014; chen et al., 2018). information about the effects of restocked eels’ density on the restocking yield are much less, and densities used for restocking are, typically, site specific and established based on the natural recruitment and yield per recruit estimates (moriarty and dekker, 1997). moreover, as few studies have contextually investigated survival, growth, dispersal, and movement of the restocked eels (shiao et al., 2006; pedersen et al., 2009; desprez et al., 2013; wickström et al., 2014; ovidio et al., 2015; sjöberg et al., 2017), little is known about the best procedure for implementing restocking with maximum survival rates in riverine ecosystems and, even, about how to accurately assess the level of restocking success (pedersen 2000; pedersen, 2009; deprez et al., 2013; matondo et al., 2019). no nco mm er cia l u se on ly c. podda et al.54 based on the above cues and considering the still large gaps of knowledge about the best protocol to restock efficiently depaupered eels’ populations (wickström and sjöberg, 2014; stacey et al., 2015), we claim the need of new and science-based assessments of restocking protocols in different scenarios, possibly coping with the expected habitat quality modifications caused by climate change and unintentionally to the anthropogenic emergence and spread of pathogens (e.g., anguillicola crassus kirk, 2003; wickström et al., 2014, and anguillid herpesvirus 1, anghv-1, kullmann et al., 2017) (delrez et al., 2021). the way forward despite the research effort to date, identification of the best technologies to reduce the threats that impair a. anguilla remains challenging. data about the distribution range of the european eel are still spatially and temporally fragmented, and the available ones are still affected by a large heterogeneity in the sampling methods and in analysis protocols. these gaps of knowledge represent altogether major biases for any possible generalization about the life cycle of eels. thus, the put in place of standardized monitoring programmes represents a priority to increase our knowledge of the eels’ life cycle and their migration patterns. only when these gaps of knowledge will be filled, restoration of environmental connectivity, particularly when rivers’ flow is interrupted by artificial obstacles like dams, will contribute to enhance eels’ stocks and their ability to fuel future generations. in this regard, we anticipate that the removal or mitigation of migration barriers, by promoting fish passage and habitat restoration, could represent a key step to enhance the yield of any eventual restocking practice without prejudice to the risk that restoring connectivity could facilitate the dispersion of alien fish species in a catchment (clavero and hermoso, 2010). better understand habitat-eel relationships is probably one of the most promising ways that may contribute to habitat restoration for restoring inland eel stocks (laffaille et al., 2004). using eels to study water contamination based on an integrated approach (ecotoxicological, parasitological, pollution topics) is crucial for the evaluation of environmental health, and chemical status of water bodies, and will directly be beneficial for restoration of eels’ stocks and consequently for ensuring water quality and habitat conservation (maes et al., 2005; belpaire and goemans, 2007; bourillon et al., 2020; capoccioni et al., 2020). with this in mind, we contend that identifying river basins and the minimum proportion of river stretches that could serve as “eel reserves” is also needed, along with collaborative research approach between researchers and stakeholders, with the final aim of establishing protocols of eels’ exploitation that respond to the principles of a sustainable use of resources and development. corresponding author: cpodda@unica.it keywords: anguilla anguilla, catadromous fish, migratory behaviour, recruitment, dams, restocking contributions: all the authors have read and approved the final version of the manuscript and agreed to be accountable for all aspects of the work. conflict of interest: the authors declare no potential conflict of interest. availability of data and materials: all data generated or analyzed during this study are included in this published article. acknowledgements: cinzia podda gratefully acknowledges sardinian regional government for the financial support of her phd scholarship (p.o.r. sardegna f.s.e. operational programme of the autonomous region of sardinia, european social fund 20142020 axis iii education and training, thematic goal 10, investment priority 10ii), specific goal 10.5. received: 22 july 2021. accepted: 18 october 2021. this work is licensed under a creative commons attribution noncommercial 4.0 license (cc by-nc 4.0). ©copyright: the author(s), 2021 licensee pagepress, italy advances in oceanography and limnology, 2021; 12:9997 doi: 10.4081/aiol.2021.9997 references aalto e, capoccioni f, terradez mas j, schiavina m, leone c, de leo g, ciccotti e, 2016. quantifying 60 years of declining european eel (anguilla anguilla l., 1758) fishery yields in mediterranean coastal lagoons. ices j. mar. sci. 73:101-110. aarestrup k, økland f, hansen mm, righton d, gargan p, castonguay m, bernatchez l, howey p, sparholt h, pedersen mi, mckinley rs, 2009. oceanic spawning migration of the european eel (anguilla anguilla). science 325:1660. acou a, 2006. [bases biologiques d’un modéle pour estimer la biomasse féconde de l’anguille europénne en fonction des recrues fluviales et du contexte de croissance: approche comparative à l’échelle de petits bassins versants, ph.d. thesis.].[article in french]. france: university of rennes 1: 315 pp. http://www.diadfish.org/francais/documents_f.htm. adam g, feunteun e, prouzet p, rigaud c, 2008. [l’anguille européenne: indicateurs d’abondance et de colonisation.].[article in french]. coll. savoir-faire, editions quae, versailles: 400 pp. alexander j, wilson r, green w, 2012. a brief history and summary of the effects of river engineering and dams on the mississippi river system and delta. circular 1; us. geological survey circular 1375: reston, virginia: 43 pp. amara r, selleslagh j, billon g, minier c, 2009. growth and condition of 0-group european flounder, platichthys flesus as indicator of estuarine habitat quality. hydrobiologia 627:87-98. no nco mm er cia l u se on ly hard times for the catadromous european eel 55 amilhat e, aarestrup k, faliex e, simon g, westerberg h, righton d, 2016. first evidence of european eels exiting the mediterranean sea during spawning migration. sci. rep. 6:21817. anderson mj, 1991. a new method for non-parametric multivariate analysis of variance. austral. ecol. 26:32-46. arai t, aoyama j, ishikawa s, miller mj, otake t, inagaki t, tsukamoto k, 2001. early life history of tropical anguilla leptocephali in the western pacific ocean. mar. biol. 138:887-895. arai t, daniel l, otake t, tsukamoto k, 1999. metamorphosis and inshore migration of tropical eels, anguilla spp., in the indo-pacific. mar. ecol. prog. ser. 182:283-293. arai t, kotake a, harrod c, morrissey m, mccarthy tk, 2019. ecological plasticity of the european eel anguilla anguilla in a tidal atlantic lake system in ireland. j. mar. biolog. assoc. u.k. 99:1189-1195. arai t, kotake a, mccarthy tk, 2006. habitat use by the european eel anguilla anguilla in irish waters. estuar. coast. shelf sci. 67:569-578. arai t, miller mj, tsukamoto k, 2003. larval duration of the tropical eel anguilla celebesensis from indonesian and philippine coasts. mar. ecol. prog. ser. 251:255-261. aranburu a, estibaliz d, briand c, 2015. glass eel recruitment and exploitation in a south european estuary (oria, bay of biscay). ices j. mar. sci. 73:111-121. arribas c, fernández-delgado c, oliva-paterna fj, drake p, 2012. oceanic and local environmental conditions as forcing mechanisms of the glass eel recruitment to the southernmost european estuary. estuar. coast. shelf sci. 107:46-57. atkinson s, carlsson jel, ball b, egan d, kelly-quinn m, whelan k, carlsson j, 2018. a quantitative pcr based environmental dna assay for detecting atlantic salmon (salmo salar l.) aquat. conserv. mar. freshwat. ecosyst. 28:1238-1243. atta ki, 2013. morphological, anatomical and histological studies on the olfactory organs and eyes of teleost fish: anguilla anguilla in relation to its feeding habits. jobaz 66:101-108. aubrun l, 1986. [inventaire de l’exploitation de l’anguille sur le littoral de la bretagne.][article in french]. ensar, rennes: 107 pp. aubrun l, 1987. [inventaire de l”exploitation de l’anguille sur le littoral sud gascogne.][article in french]. laboratoire de biologie halieutique, ensa rennes: 150 pp. baltazar-soares m, biastoch a, harrod c, hanel r, marohn l, prigge e, evans d, bodles k, behrens e, w. böning k, eizaguirre c, 2014. recruitment collapse and population structure of the european eel shaped by local ocean current dynamics. curr. biol. 24:104-108. baltazar-soares m, eizaguirre c, 2017. animal navigation: the eel’s magnetic guide to the gulf stream. curr. biol. 27:604-606. barry j, newton m, dodd ja, hooker oe, boylan p, lucas mc, adams ce, 2015. foraging specialisms influence space use and movement patterns of the european eel anguilla anguilla. hydrobiologia 766:333-348. beaulaton l, castelnaud g, 2005. the efficiency of selective tidal stream transport in glass eel entering the gironde (france). bull. fr. pêche piscic. 5:378-379. beaulaton l, drouineau, h, 2018. modelling the recruitment of european eel (anguilla anguilla) throughout its european range. ices j. of mar. sci. 75:541-552. béguer-pon m, castonguay m, shan s, benchetrit j, dodson jj, 2015. direct observations of american eels migrating across the continental shelf to the sargasso sea. nat. commun. 6:8705. béguer-pon m, dodson jj, castonguay m, jellyman d, aarestrup k, tsukamoto k, 2018. tracking anguillid eels: five decades of telemetry-based research. mar. freshw. res. 69:199. béguer-pon m, ohashi k, sheng j, castonguay m, dodson jj, 2016. modeling the migration of the american eel in the gulf of st. lawrence. mar. ecol. prog. ser. 549:183-198. behrmann-godel j, eckmann r, 2003. a preliminary telemetry study of the migration of silver european eel (anguilla anguilla l.) in the river mosel, germany ecol. freshw. fish 12:196-202. belpaire c, goemans g, 2007. the european eel anguilla anguilla, a rapporteur of the chemical status for the water framework directive? vie et milieu life and environment. 57: 235-252. benchetrit j, béguer-pon m, sirois p, castonguay m, fitzsimons j, dodson jj, 2017. using otolith microchemistry to reconstruct habitat use of american eels anguilla rostrata in the st. lawrence river-lake ontario system. ecol. freshw. fish 26:19-33. benson t, de bie j, gaskell j, vezza p, kerr jr, lumbroso d, owen mr, kemp ps, 2021. agent-based modelling of juvenile eel migration via selective tidal stream transport. ecological modelling 443:109448. bertin l, 1956. eels, a biological study. cleaver-hume press ltd, london. besson ml, trancart t, acou a, charrier f, mazel v, legault a, feunteun e, 2016. disrupted downstream migration behaviour of european silver eels (anguilla anguilla, l.) in an obstructed river. environ. biol. fishes 99:779-791. bevacqua d, melià p, gatto m, de leo ga, 2015. a global viability assessment of the european eel. glob. chang. biol. 21:3323-3335. bevacqua d, melià p, schiavina m, crivelli aj, de leo ga, gatto m, 2019. a demographic model for the conservation and management of the european eel: an application to a mediterranean coastal lagoon. ices j. mar. sci. 76: 2164-2178. birnie-gauvin k, tummers js, lucas mc, aarestrup k, 2017. adaptivemanagement in the context of barriers in european freshwater ecosystems. j. environ. manag. 204:436-441. bisgaard j, pedersen mi, 1991. mortality and growth of wild and introduced cultured eels (anguilla anguilla (l)) in a danish stream, with special reference to a new tagging technique. dana 9:57-69. boëtius j, harding ef, 1985. a re-examination of johannes schmidt’s atlantic eel investigations. dana 4:129-162. bonhommeau s, castonguay m, rivot e, sabatié r, le pape o, 2010. the duration of migration of atlantic anguilla larvae. fish fish. 11:289-306. bonhommeau s, chassot e, rivot e, 2008. fluctuations in european eel (anguilla anguilla) recruitment resulting from environmental changes in the sargasso sea. fish. oceanogr. 17:32-44. no nco mm er cia l u se on ly c. podda et al.56 bornarel v, lambert p, briand c, antunes c, belpaire c, ciccotti e, diaz e, diserud o, doherty d, domingos i, evans d, de graaf m, o’leary c, pedersen m, poole r, walker a, wickström h, bouvet j, prouzet p, bru n, de casamajor m, lissardy m, sanchez f, 2006. [quantification de la biomasse saisoniére de civelles (anguilla anguilla) dans l’estuaire de l’adour et estimation du taux d’exploitationsaisonnier de la pêche professionnelle au tamis poussé.].[article in french]. technical report, ifremer. http://www.ifremer.fr/indicang/ sites-thematiques/pdf/flux-saison.pdf. boubee ja, mitchell cp, chisnall bl, west dw, bowman ej, haro a, 2001. factors regulating the downstream migration of mature eels (anguilla spp,) at aniwhenua dam, bay of plenty, n. z. j. mar. freshwater res. 35:121-134. bourillon b, acou a, trancart t, belpaire c, covaci a, bustamante p, faliex e, amilhat e, malarvannan g, virag l, aarestrup k, bervoets l, boisneau c, boulenger c, gargan p, becerra-jurado g, lobón-cerviá j, maes ge, pedersen mi, poole r, sjöberg n, wickström h, walker a, righton d, feunteun e, 2020. assessment of the quality of european silver eels and tentative approach to trace the origin of contaminants a european overview. sci. total environ. 743:140675. bradford rg, carr jw, page fh, whoriskey f, 2009. migration of silver american eels through a macrotidal estuary and bay, p.275-292. in: challenges for diadromous fishes in a dynamic global environment. haro a, smith kl, rulifson ra, moffitt cm, klauda rj, dadswell mj, cunjak ra, cooper je, beal kl, avery ts (eds.). american fisheries society symposium, bethesda, md. bracken fsa, rooney sm, kelly-quinn m, king jj, carlsson j, 2019. identifying spawning sites and other critical habitat in lotic systems using edna “snapshots”: a case study using the sea lamprey petromyzon marinus l. ecol. evol. 9:553-567. breukelaar aw, ingendahl d, vriese ft, de laak g, staas s, klein breteler jgp, 2009. route choices, migration speeds and daily migration activity of european silver eels anguilla anguilla in the river rhine, north-west europe. j. fish. biol. 74:2139-2157. brink k, gough p, royte j, schollema pp, wanningen h, 2018. ‘from sea to source 2.0. protection and restoration of fish migration in rivers worldwide.’ (world fish migration foundation: groningen, netherlands.): 364 pp. bru n, prouzet p, lejeune m, 2009. daily and seasonal estimates of the recruitment and biomass of glass eels runs (anguilla anguilla) and exploitation rates in the adour open estuary (southwestern france). aquatic living resources 22:509-523. bruijs m, durif c, 2009. silver eel migration and behaviour, p.65-95. in: thillart g, dufour s, rankin jc (eds.), spawning migration of the european eel. bunn se, arthington ah, 2002. basic principles and ecological consequences of altered flow regimes for aquatic biodiversity. environ. manage. 30:492-507. bureau du colombier s, bolliet v, lambert p, bardonnet a, 2007. energy and migratory behaviour in glass eels (anguilla anguilla). physiol. behav. 92:684-690. burkhead nm, 2012. extinction rates in north american freshwater fishes, 1900–2010. bioscience 62:798-808. byrne d, beckett, b, 2012.wicklow bridges project. assessment of the risk of barriers to migration of fish species in county wicklow. retrieved from. http://www.countywicklowheritage.org/documents/wicklow_bridges_project_fish _passage_report.pdf. calles o, olsson ic, comoglio c, kemp ps, blunden l, schmitz m, greenberg la, 2010. size-dependent mortality of migratory silver eels at a hydropower plant, and implications for escapement to the sea. freshw. biol. 55:2167-2180. capoccioni f, leone c, belpaire c, malarvannan g, poma g, de matteis g, tancioni l, contò m, failla s, covaci a, ciccotti e, 2020. quality assessment of escaping silver eel (anguilla anguilla l.) to support management and conservation strategies in mediterranean coastal lagoons. environ. monit. assess. 192:570. carlsson j, aarestrup k, nordwall f, näslund i, eriksson t, carlsson jel, 2004. migration of landlocked brown trout in two scandinavian streams as revealed from trap data. ecol. freshw. fish 13:161-167. carvajal-quintero, jd, januchowski-hartley sr, maldonadoocampo ja, jézéquel c, delgado j, tedesco pa, 2017. damming fragments species ranges and heightens extinction risk. conserv. lett. 10:708-716. castonguay m, dutil jd, audet c, miller r, 1990. locomotor activity and concentration of thyroid hormones in migratory and sedentary juvenile american eels. t. am. fish. soc. 6:946-956. chapman bb, skov c, hulthen k, brodersen j, nilsson pa, hansson la, brönmark c, 2012. partial migration in fishes: definitions, methodologies and taxonomic distribution. j. fish. biol. 81:479-499. chasco be, kaplan ic, thomas ac, acevedo-gutiérrez a, noren dp, ford mj, bradley hanson m, scordino jj, jeffries sj, marshall kn, shelton ao, matkin c, burke bj, ward ej, 2017. competing tradeoffs between increasing marine mammal predation and fisheries harvest of chinook salmon. sci. rep. 7:15439. cheng pw, tzeng wn, 1996. timing of metamorphosis and estuarine arrival across the dispersal range of the japanese eel anguilla japonica. mar. ecol. progr. ser. 131:87-96. ciccotti e, 2005. interactions between capture fisheries and aquaculture: the case of the eel (anguilla anguilla l., 1758). in: s. cataudella, d. crosetti, & f. massa (eds.), interactions between capture fisheries and aquaculture: a methodological perspective. rome: gfcm studies and reviews 78:190-203. ciccotti e, busilacchi s, cataudella s, 2000. eel, anguilla anguilla (l.), in italy: recruitment, fisheries and aquaculture. dana 12:7-15. cites (2020a). conservation status of anguilla anguilla. https://www.speciesplus. net/#/taxon_conce pts/3973/legal. cites (2020b). convention of migratory species (cms) listing of anguilla anguilla. https://www.speciesplus.net/#/taxon_ conce pts/66526/legal. clavero m, hermoso v, 2010. reservoirs promote the taxonomic homogenization of fish communities within river basins. biodivers. conserv. 20:41-57. clément m, chiasson ag, veinott g, cairns dk, 2014. what otolith microchemistry and stable isotope analysis reveal and conceal about anguillid eel movements across salinity boundaries. oecologia 175:1143-1153. no nco mm er cia l u se on ly hard times for the catadromous european eel 57 close da, fitzpatrick ms, li hw, 2002. the ecological and cultural importance of a species at risk of extinction, pacific lamprey. fisheries 27:19-25. collas fp, buijse ad, van den heuvel l, van kessel n, schoor mm, eerden h, leuven rs, 2018. longitudinal training dams mitigate effects of shipping on environmental conditions and fish density in the littoral zones of the river rhine. sci. total environ. 619-620:1183-1193. comte l, hugueny b, grenouillet g, 2016. climate interacts with anthropogenic drivers to determine extirpation dynamics. ecography 39:1008-1016. cooke sj, midwood jd, thiem jd, klimley p, lucas mc, thorstad eb, eiler j, holbrook c, ebner bc, 2013. tracking animals in freshwater with electronic tags: past, present and future. anim. biotelemetry 1:5-19. costa-dias s, sousa r, lobón-cerviá j, laffaille p, 2009. the decline of diadromous fish in western europe inland waters: mains causes and consequence, p.67-92. in: fisheries: management, economics and perspectives. nova science publishers. cresci a, 2020. a comprehensive hypothesis on the migration of european glass eels (anguilla anguilla). biol. rev. 95:1273-1286. cresci a, durif cm, paris cb, shema sd, skiftesvik ab, browman hi, 2019b. glass eels (anguilla anguilla) imprint the magnetic direction of tidal currents from their juvenile estuaries. commun. biol. 2:366. cresci a, paris cb, durif cm, shema s, bjelland rm, ab skiftesvik, browman hi, 2017. glass eels (anguilla anguilla) have a magnetic compass linked to the tidal cycle sci. adv. 3:e1602007. cresci a, paris cb, foretich ma, durif cm, shema sd, o’brien ce, vikebø fb, skiftesvik ab, browman hi, 2019a. atlantic haddock (melanogrammus aeglefinus) larvae have a magnetic compass that guides their orientation. iscience 19:1173-1178. crivelli a, auphan n, chauvelon p, sanzoz a, menella j, poizat g, 2008. glass eel recruitment, anguilla anguilla (l.), in a mediterranean lagoon assessed by a glass eel trap: factors explaining the catches. hydrobiologia 602:79-86. crnjar r, slcalera g, bigiani a, tomassini barbarossa i, magherini pc, pietra p, 1992. olfactory sensitivity to amino acids in the juvenile stages of the european eel anguilla anguilla (l.). j. fish biol. 40:567-576. culurgioni j, figus v, cabiddu s, de murtas r, cau a, sabatini a, 2015. larval helminth parasites of fishes and shellfishes from santa gilla lagoon (sardinia, western mediterranean), and their use as bioecological indicators. estuaries coasts 38:1505-1519. culurgioni j, sabatini a, de murtas r, mattiucci s, figus v, 2014. helminth parasites of fish and shellfish from the santa gilla lagoon in southern sardinia, italy. j. helminthol 88:489-498. cushing dh, 1990. plankton production and year-class strength in fish populations: an update of the match/mismatch hypothesis adv. mar. biol. 26:249-293. cullen p, mccarthy tk, 2003. hydrometric and meteorological factors affecting the seaward migration of silver eels (anguilla anguilla, l.) in the lower river shannon. environ. biol. fishes 67:349-357. de casamajor mn, bru n, prouzet p, 1999. [influence de la luminosité nocturne et de la turbidité sur le comportement vertical de migration de la civelle d’anguille (anguilla anguilla l.) dans l’estuaire de l’adour.].[article in french] bull. fr. pêche piscic. 355:327-347. de leo ga, melià paco, gatto m, crivelli aj, 2009. eel population modeling and its application to conservation management. in: casselman jm and cairns dk (eds.), eels at the edge: science, status, and conservation concerns. proceedings of the 2003 international eel symposium. am. fish soc. symp. 58:327-345. deiner k, walser jc, mächler e, altermatt f, 2015. choice of capture and extraction methods affect detection of freshwater biodiversity from environmental dna. biol. conserv. 183:53-63. dejean t, valentini a, duparc a, pellier-cuit s, pompanon f, taberlet p, miaud c, 2011. persistence of environmental dna in freshwater ecosystems. plos one 6:8-11. delpech c, courrat a, pasquaud s, lobry j, le pape o, nicolas d, boët p, girardin m, lepage m, 2010. development of a fish-based index to assess the ecological quality of transitional waters: the case of french estuaries. mar. pollut. bull. 60:908-918. dekker w, 1998. long-term trends in the glass eels immigrating at den oever, the netherlands. bull. fr. pêche piscic. 349:199-214. dekker w, 2000. the fractal geometry of the european eel stock. ices j. mar. sci. 57:109-121. dekker w, 2003a. did lack of spawners cause the collapse of the european eel, anguilla anguilla? fish. manag. ecol. 10:365-376. dekker w, 2003b. status of the european eel stock and fisheries, p.237-254. in: aida k., tsukamoto k., yamauchi k. (eds), eel biology. springer, tokyo. dekker w, 2004. slipping through our hands – population dynamics of the european eel. (phd thesis). amsterdam, the netherlands: university of amsterdam: 186 pp. http://www.diadfish.org/doc/these_2004/dekker_thesis_eel. pdf.dekker w, casselman jm, 2014. dekker w, 2016. management of the eel is slipping through our hands! distribute control and orchestrate national protection. ices j. mar. sci. 73:2442-2452. dekker w, 2019. the history of commercial fisheries for european eel commenced only a century ago. fish. manag. ecol. 6:6-19. dekker w, beaulaton l, 2016a. climbing back up what slippery slope? dynamics of the european eel stock ad its management in historical perspective. ices j. mar. sci. 73:5-13. dekker w, beaulaton l, 2016b. faire mieux que la nature the history of eel restocking in europe. environ. hist. 22: 255-300. dekker w, casselman jm, 2014. the 2003 québec declaration of concern about eel declines 11 years later: are eels climbing back up the slippery slope? fisheries 39:613-614. dekker w, casselman jm, cairns dk, tsukamoto k, jellyman d, lickers h, 2003. worldwide decline of eel resources necessitates immediate action: québec declaration of concern. fisheries 28:28-30. desaunay y, guerault d, 1997. seasonal and long-term changes in biometrics of eel larvae: a possible relationship between no nco mm er cia l u se on ly c. podda et al.58 recruitment variation and north atlantic ecosystem productivity. j. fish biol. 51:317-339. delrez n, zhang h, lieffrig f, mélard c, farnir f, boutier m, donohoe o, vanderplasschen a, 2021. european eel restocking programs based on wild-caught glass eels: feasibility of quarantine stage compatible with implementation of prophylactic measures prior to scheduled reintroduction to the wild. j. nat. conserv. 59:125933. desprez m, crivelli aj, lebel i, massez g, gimenez o, 2013. demographic assessment of a stockingexperiment in european eels. ecol. freshw. fish 22:412-420 dias ms, tedesco pa, hugueny b, jézéquel c, beauchard o, brosse s, oberdorff t, 2017. anthropogenic stressors and riverine fish extinctions. ecol. indic. 79:37-46. domingos i, 1992. fluctuation of glass eel migration in the mondego estuary (portugal) in 1988 and 1989. irish fisheries investigations: series a (freshwater) (ireland) 36:1-4. dorow m, schulz s, frankowski j, ubl c, 2019. using a telemetry study to assess the boundary net efficiency of an enclosure system used for yellow eel density monitoring. fish. manag. ecol. 26:70-75. drouineau h, bau f, alric a, deligne n, gomes p, sagnes p, 2017. silver eel downstream migration in fragmented rivers: use of a bayesian model to track movements triggering and duration aquat. living resour. 30:1-9. drouineau h, briand c, lambert p, beaulaton p, 2018. gerem (glass eel recruitment estimation model): a model to estimate glass eel recruitment at different spatial scales. fish res. 174:68-80. drouineau h, carter c, rambonilaza m, beaufaron g, bouleau g, gassiat a, lambert p, le floch s, tétard s, de oliveira e, 2018b. river continuity restoration and diadromous fishes: much more than an ecological issue. environ. manage. 61:671-686. drouineau h, durif c, castonguay m, mateo m, rochard e, verreault g, yokouchi k, lambert p, 2018a. freshwater eels: a symbol of the effects of global change. fish fish. 19:903-930. dudgeon d, arthington ah, gessner mo, kawabata zi, knowler dj, lévêque c, naiman rj, prieur-richard ah, soto d, stiassny mlj, sullivan ca, 2006. freshwater biodiversity: importance, threats, status and conservation challenges. biol. rev. 81:163-182. durant jm, hjermann do, ottersen g, stenseth nc, 2007. climate and the match or mismatch between predator requirements and resource availability clim. res. 33:271-283. durif cmf, 2003. [la migration d’avalaison de l’anguille européenne anguilla anguilla: caractérisation des fractions dévalantes, phénoméne de migration et franchissement d’obstacles.][article in french]. thèse de doctorat, université paul sabatier, toulouse: 350 pp. durif cmf, elie p, 2008. predicting downstream migration of silver eels in a large river catchment based on commercial fishery data. fish manag. ecol. 15:127-137. durif cmf, gjosaeter j, vollestad la, 2011. influence of oceanic factors on anguilla anguilla (l.) over the twentieth century in coastal habitats of the skagerrak, southern norway. proc. r. soc. lond. series b 278:464-473. edeline e, beaulaton l, le barh r, elie p, 2007. dispersal in metamorphosing anguilla anguilla juvenile eel. mar. ecol. prog. ser. 344:213-218. edeline e, dufour s, briand c, fatin d, elie p, 2004. thyroid status is related to migratory behavior in anguilla anguilla glass eels. mar. ecol. prog. ser. 282:161-270. edeline e, dufour s, elie p, 2005. role of glass eel salinity preference in the control of habitat selection and growth plasticity in anguilla anguilla. mar. ecol. prog. ser. 304:191-199. edeline e, dufour s, elie p (2009) proximate and ultimate control of eel continental dispersal. in: van den thillart g, dufour s, rankin jc (eds.) spawning migration of the european eel: reproduction index, a useful tool for conservation management. fish fish. ser. 30:433-461. eldrogi n, luthon f, larroque b, alqaddafi s, bolliet v, 2018. motion estimation of glass eels by differential methods. istj: pp.299-315. elie p, rochard e, 1994. [migration des civelles d’anguilles (anguilla anguilla l.) dans les estuaires, modalités du phénomène et caractéristiques des individus.][article in french]. bull. fr. pêche piscic. 335:81-98. elliott m, whitfield ak, potter ic, blaber sjm, cyrus dp, nordlie fg, harrison td, 2007. the guild approach to categorizing estuarine fish assemblages: a global review. fish fish. 8:241-268. enbody ed, pettersson me, sprehn cg, palm s, wickströmb h, andersson l, 2021. ecological adaptation in european eels is based on phenotypic plasticity. pnas 118:e2022620118. eu (2007). council regulation (ec) no. 1100/2007 of 18 september 2007, establishing measures for the recovery of the stock of european eel. official journal of the european union l. 248:17-23. evans nt, li y, renshaw ma, olds bp, deiner k, turner cr, jerde cr, lodge dm, lamberti ga, pfrender me, 2016. fish community assessment with edna metabarcoding: effects of sampling design and bioinformatic filtering. can. j. fish. aquat. sci. 74:1362-1374. fao, 2003. inland fisheries. fao, 2007. report of the second fao ad hoc expert advisory panel for the assessment of proposals to amend appendices i and ii of cites concerning commercially exploited aquatic species. rome, 26–30 march 2007. fao fisheries report. no. 833. rome: 133 pp. feunteun e, 2002. management and restoration of the european eel population (anguilla anguilla): an impossible bargain? ecol. eng. 18:575-591. feunteun e, acou a, laffaille p, legault a, 2000. european eel (anguilla anguilla): prediction of spawner escapement from continental population parameters. can. j. fish. aquat. sci. 57:1627-1635. feunteun e, laffaille p, 2011. commercially important catadromous fish. encyclopedia of life support systems (eolss): 27 pp. feunteun e, laffaille p, robinet t, briand c, baisez a, olivier jm, acou a, 2003. a review of upstream migration and movements in inland waters by anguillid eels: towards a general theory, p.181-190. in: aida k, tsukamoto k, yamauchi k (eds.) eel biology. springer, tokyo. fischer es, blackburn se, liss sa, hughes js, li h, deng zd, 2019. how small can we go? evaluating survival, tag retention, and growth of juvenile chinook salmon imno nco mm er cia l u se on ly hard times for the catadromous european eel 59 planted with a new acoustic microtag. n. am. j. fish. manag. 39:1329-1336. friedland kd, miller mj, knights b, 2007. oceanic changes in the sargasso sea and declines in recruitment of the european eel. ices j. mar. sci. 64:519-530. forward rb, tankersley ra, 2001. selective tidal-stream transport of marine animals oceanogr. mar. biol.: ann. rev. 39:201-213. galbraith hs, blakeslee cj, schmucker ak, johnson ns, hansen mj, li w, 2017. donor life stage influences juvenile american eel anguilla rostrata attraction to conspecific chemical cues. j. fish biol. 90:384-395. gandolfi g, pesaro m, tongiorgi p, 1984. environmental factors affecting the ascent of elvers, anguilla anguilla (l.), into the arno river. oebalia 126:17-35. geffroy b, bardonnet a, 2015. sex differentiation and sex determination in eels: consequences for management. fish fish. 17:375-398. garcia de leániz, c, 2008. weir removal in salmonid streams: implications, challenges and practicalities. hydrobiologia 609:83-96. garibaldi a, turner n, 2004. cultural keystone species: implications for ecological conservation and restoration. ecol. soc. 9:1. garman gc, 1992. fate and potential significance of postspawning anadromous fish carcasses in an atlantic coastal river. trans. am. fish. soc. 121:390-394. gascuel d, feunteun e, fontenelle g, 1995. seasonal dynamics of estuarine migration in glass eels (anguilla anguilla). aquat. living resour. 8:123-133. grill g, lehner b, thieme m, geenen b, tickner d, antonelli f, babu s, borrelli p, cheng l, crochetiere h, ehalt macedo h, filgueiras r, goichot m, higgins j, hogan z, lip b, mcclain me, meng j, mulligan m, nilsson c, olden jd, opperman jj, petry p, reidy liermann c, saénz l, salinas-rodrìguez s, schelle p, schmitt rjp, snider j, tan f, tockner k, valdujo ph, van soesbergen, zarfl c, 2019. mapping the world’s free-flowing rivers. nature. 569:215-221. guarniero i, cariani a, ferrari a, sulliotti v, emmanuele p, casalini a, tinti f, mordenti o, 2020. sexual behaviour and reproductive performance of the endangered european eel anguilla anguilla (linnaeus, 1758) based on direct observations and paternity assignment in semi-natural conditions. aquac. rep. 16:100258. gutiérrez-estrada jc, pulido-calvo i, 2015. is the atlantic surface temperature a good proxy for forecasting the recruitment of european eel in the guadalquivir estuary? prog. oceanogr. 130:112-124. han ys, hsiung km, zhang h, chow ly, tzeng wn, shinoda a, yoshinaga t, hur sp, hwang sd, iizuka y, kimura s, 2019. dispersal characteristics and pathways of japanese glass eel in the east asian continental shelf. sustainability 11:2572. han ys, lin yf, wu cr, iizuka y, castillo tr, yambot iu, mamalangkap md, yambot av, 2016. biogeographic distribution of the eel anguilla luzonensis: dependence upon larval duration and oceanic currents. mar. ecol. prog. ser. 551:227-238. hanel r, briand c, diaz e, döring r, sapounidis a, warmerdam w, andrés m, freese m, marcelis a, marohn l, pohlmann jd, van scharrenburg m, waidmann n, walstra j, werkman m, de wilde j, wysujack k, 2019. research for pech committee – environmental, social and economic sustainability of european eel management. european parliament, policy department for structural and cohesion policies, brussel: 192 pp. hanel r, stepputtis d, bonhommeau s, castonguay m., schaber m., wysujack k, vobach m, miller mj, 2014. low larval abundance in the sargasso sea: new evidence about reduced recruitment of the atlantic eels. naturwissenschaften 101:1041-1054. harrison aj, walker am, pinder ac, briand c, aprahamian mw, 2014. a review of glass eel migratory behaviour, sampling techniques and abundance estimates in estuaries: implications for assessing recruitment, local production and exploitation. rev. fish biol. fish. 24:967-983. hewavitharane ca, pickering td, rico c, mochioka n, 2020. early life history of tropical freshwater eels (anguilla spp.) recruiting to viti levu, fiji islands, in the western south pacific. mar. freshw. res. 71:452-460. higuchi t, watanabe s, manabe r, kaku t, okamura a, yamada y, miller mj, tsukamoto k, 2018. tracking anguilla japonica silver eels along the west marina ridge using popup archival transmitting tags. zool. stud. 57:24. huertas m, canàrio avm, hubbard pc, 2008. chemical communication in the genus anguilla: a minireview. behaviour 145:1389-1407. ices, 2011. report of the 2011 session of the joint eifaac/ices working group on eels (wgeel), 5-9 september 2011, lisbon, portugal: 244 pp. ices, 2016. report of the working group on eels (wgeel), 1522 september 2016, cordoba, spain. (ices cm 2016/acom: 19). http://ices.dk/sites/pub/publication%20reports/expert%20group%20report/acom/2016/wgeel/wge el_2016.pdf. ices, 2018. report of the joint eifaac/ices/gfcm working group on eels (wgeel), 5-12 september 2018, gdańsk, poland. ices cm 2018/acom:15, 152 pp. ices. 2019. report of the joint eifaac/ices/gfcm working group on eels (wgeel), ices scientific reports. 1:50, 177 pp. ices, 2020. report of the joint eifaac/ices/gfcm working group on eels (wgeel). ices scientific reports. 2:85, 223 pp. imbert h, labonne j, rigaud c, lambert p, 2010. resident and migratory tactics in freshwater european eels are sizedependent. freshw. biol. 55:1483-1493. iucn, 2014. the iucn red list of threatened species: anguilla anguilla. http://www.iucnredlist.org. iucn, 2019. the iucn red list of threatened species. http://www.iucnredlist.org. jacoby dmp, casselman jm, crook v, delucia mb, ahn h, kaifu k, kurwie t, sasal p, silfvergrip amc, smith kg, uchida k, walker am, gollock mj, 2015. synergistic patterns of threat and the challenges facing global anguillid eel conservation. gecco 4:321-333. januchowski-hartley sr, mcintyre pb, diebel m, doran pj, infante dm, joseph c, allan jd, 2013. restoring aquatic ecosystem connectivity requires expanding inventories of no nco mm er cia l u se on ly c. podda et al.60 both dams and road crossings. front. ecol. environ. 11:211-217. jegstrup im, rosenkilde p, regulation of post-larval development in the european eel: thyroid hormone level, progress of pigmentation and changes in behaviour j. fish biol. 63:168-175. jellyman dj, arai t, 2016. juvenile eels: upstream migration and habitat use, p.171-191. in: arai t, (ed.), biology and ecology of anguillid eels. jellyman dj, glova gj, todd pr, 1996. movements of shortfinned eels, anguilla australis, in lake ellesmere, new zealand: results from mark-recapture studies and sonic tracking. n.z. j. mar. freshw. res. 30:371-381. jellyman dj, lambert pw, 2003. factors affecting recruitment of glass eels into the grey river, new zealand. fish biol. 63:1067-1079. jellyman dj, tsukamoto k, 2002. first use of archival transmitters to track migrating freshwater eels anguilla dieffenbachii at sea. mar. ecol. prog. ser. 233:207-215. jones j, börger l, tummers j, jones p, lucas m, kerr j, kemp p, bizzi s, consuegra s, marcello l, vowles a, belletti b, verspoor e, bund de gough, p, garcia de leániz c, 2019. a comprehensive assessment of stream fragmentation in great britain. sci. total environ. 673:756-762. katz j, moyle pb, quinous rm, israel j, purdy s, 2013. impending extinction of salmon, steelhead, and trout (salmonidae) in california. environ. biol. fish. 96:1169-1186. kemp ps, o’hanley jr, 2010. procedures for evaluating and prioritising the removal of fish passage barriers: a synthesis. fish. manag. ecol. 17:297-322. kettle aj, bakker dce, haines k, 2008. impact of the north atlantic oscillation on the trans-atlantic migrations of the european eel (anguilla anguilla). j. geophys. res. 113:26. kettle aj, haines k, 2006. how does the european eel (anguilla anguilla) retain its population structure during its larval migration across the north atlantic ocean? can. j. fish. aquat. sci. 63:90-106. kettle aj, vøllestad la, wibig j, 2011. where once the eel and the elephant were together: decline of the european eel because of changing hydrology in southwest europe and northwest africa? fish fish. 12:380-411. kirk rs, 2003. the impact of anguillicola crassus on european eels. fish. manag. ecol. 10:385-394. knights b, 2003. a review of the possible impacts of long-term oceanic and climate changes and fishing mortality on recruitment of anguillid eels of the northern hemisphere. sci. total environ. 310:237-244. knudsen sw, ebert rb, hesselsøe m, kuntke f, hassingboe j, bondgaard mortensen p, thomsen pf, sigsgaard ee, klitgaard hansen b, nielsen eeg, møller pr, 2019. speciesspecific detection and quantification of environmental dna from marine fishes in the baltic sea. j. exp. mar. biol. ecol. 510:31-45. kroon fj, phillips s, 2016. identification of human-made physical barriers to fish passage in the wet tropics region, australia. mar. freshw. res. 67:677-681. kullmann b, adamek m, steinhagen d, thiel r, 2017. anthropogenic spreading of anguillid herpesvirus 1 by stocking of infected farmed european eels, anguilla anguilla (l.), in the schlei fjord in northern germany. j. fish dis. 40:1695-1706. kullmann b, thiel r, 2018. bigger is better in eel stocking measures? comparison of growth performance, body condition, and benefit-cost ratio of simultaneously stocked glass and farmed eels in a brackish fjord. fish. res. 205:132-140. laffaille p, baisez a, rigaud c, feunteun e, 2004. habitat preferences of different european eel size classes in a reclaimed marsh: a contribution to species and ecosystem conservation. wetlands 24:642-651. laffaille p, caraguel jm, legault a, 2007. temporal patterns in the upstream migration of european glass eels (anguilla anguilla) at the couesnon estuarine dam. estuar. coast. shelf sci. 73:81-90. laffaille p, feunteun e, baisez a, robinet t, acou t, legault a, lek s, 2003. spatial organisation of european eel (anguilla anguilla l.) in a small catchment. ecol. freshw. fish 12:254-264. laffaille p, rigaud c, 2008. [indicateurs de colonisation et de sédentarisation, p.230-275.][article in french]. in: adam g, feunteun e, prouzet p, rigaud c (eds.), l’anguille européenne, indicateurs d’abondance et de colonisation, quae, paris. lagarde r, peyre j, amilhat e, bourrin f, prellwitz f, simon g, faliex e, 2021. movements of non-migrant european eels in an urbanised channel linking a mediterranean lagoon to the sea. water 13:839. lambert p, lassalle g, acolas ml, bau f, castelnaud g, daverat f, jatteau p, rigaud c, rochard e, roqueplo c, de jouvenel f, 2019. a foresight analysis in fisheries science: the case study of migratory fish research. futures 111:90-103. larinier m, 2000. dams in fish migration, p.1-23. in: berkamp mccartney m, dugan p, mcneely j, acreman mg (eds.). dams, ecosystem functions and environmental restoration. cape town. larinier m, 2001. enviromental issues, dams and fish migration, p.45-90. in: marmulla, g. (ed.), dams, fish and fisheries. opportunities, challenges and conflict resolution. fao fisheries technical paper. no. 419. fao, rome. larinier, m., 2002. location of fishways. bull. fr. pêche piscic. 364:39-53 larinier m, travade f, 2002. downstream migration: problems and facilities. bull. fr. pêche piscic. 364:181-207. lasne e, laffaille p, 2009. assessing the freshwater distribution of yellow eel. knowl. managt. aquatic ecosyst. 04:390-391. lassalle g, rochard e, 2009. impact of twenty-first century climate change on diadromous fish spread over europe, north africa and the middle east. glob. chang. biol. 15:1072-1089. lecomte-finiger r, 1992. growth history and age at recruitment of european glass eels (anguilla anguilla) as revealed by otolith microstructure. mar. biol. 114:205-210. lin yj, lozys l, shiao jc, lizukaand y, tzeng wn, 2007. growth differences between naturally recruited and stocked european eel anguilla anguilla from different habitats in lithuania. j. fish biol. 71:1773-1787. lin yj, yalçin-özdilek s, iizuka y, gümüş a, tzeng wn, 2011. migratory life history of european eel anguilla anguilla from freshwater regions of the river asi, southern turkey and their high otolith sr:ca ratios. j. fish biol. 78:860-868. liss sa, znotinas kr, blackburn se, fischer es, hughes js, harnish ra, li h, deng zd, 2021. from 95 to 59 millimeno nco mm er cia l u se on ly hard times for the catadromous european eel 61 tres: a new active acoustic tag size guideline for salmon. can. j. fish. aquat. sci. 78:943-957. lucas mc, baras e, 2001. migration of freshwater fishes. blackwell science ltd, oxford: 412 pp. maitland ps, 2003. ecology of the river, brook and sea lamprey. conserving natura 2000. rivers ecology series no. 5. english nature, peterborough: 54 pp. maes ge, raeymaekers jam, pampoulie c, seynaeve a, goemans g, belpaire c, volckaert fam, 2005. the catadromous european eel anguilla anguilla (l.) as a model for freshwater evolutionary ecotoxicology: relationship between heavy metal bioaccumulation, condition and genetic variability. aquat. toxicol. 73:99-114. manabe r, aoyama j, watanabe k, kawai m, miller mj, tsukamoto k, 2011. first observations of the oceanic migration of japanese eel, from pop-up archival transmitting tags. mar. ecol. prog. ser. 437:229-240. marui m, arai t, miller mj, jellyman dj, tsukamoto k, 2001. comparison of early life history between new zealand temperate eels and pacific tropical eels revealed by otolith microstructure and microchemistry. mar. ecol. prog. ser. 213:273-284. matondo nb, benitez jp, dierckx a, rollin x, ovidio m, 2020. an evaluation of restocking practice and demographic stock assessment methods for cryptic juvenile european eel in upland rivers. sustainability 12:1124. matondo nb, séleck e, dierckx a, benitez jp, rollin x, ovidio m, 2019. what happens to glass eels after restocking in upland rivers? a long-term study on their dispersal and behavioural traits. aquat. conserv. mar. freshw. ecosyst. 29:374-388. mccleave jd, 2001. fish: eels, p.800-809. in: steele j. and mcneil c. (eds.). encyclopedia of ocean sciences. academic press, london. mccleave jd, kleckner rc, 1985. oceanic migrations of atlantic eels (anguilla spp.): adults and their offspring. contr. mar. sci. 27:316-337. mcdowall rm, 1988. diadromy in fishes: migration between freshwater and marine environments. croom helm, london: 308 pp. mcdowall rm, 1992. diadromy-origins and definitions of terminology. copeia: 248-251. mcintyre pb, reidy liermann c, childress e, hamann ej, hogan jd, januchowski-hartley sr, koning aa, neesontm, oele dl, pracheil bm, 2016. conservation of migratory fishes in freshwater ecosystems, p.324-360. in: conservation of freshwater fishes, closs gp, krkosek m, olden jd (eds). cambridge university press: cambridge. meybeck m, 2003. global analysis of river systems: from earth system controls to anthropocene syndromes. philos. trans. r. soc., series b, biological sciences 358:1935-1955. milardi m, lanzoni m, gavioli a, fano ea, castaldelli g, 2018. tides and moon drive fish movements in a brackish lagoon. estuar. coast. shelf sci. 215:207-214. miller mj, bonhommeau s, munk p, castonguay m, hanel r, mccleave jd, 2015. a century of research on the larval distributions of the atlantic eels: a re-examination of the data. biol. rev. 90:1035-1064. miller mj, feunteun e, tzukamoto k, 2016. did a “perfect storm” of oceanic changes and continental anthropogenic impacts cause northern hemisphere anguillid recruitment reductions? ices j. mar. sci. 73:43-56. miller mj, kimura s, friedland kd, knights b, kim h, jellyman dj, tsukamoto k, 2009. review of ocean-atmospheric factors in the atlantic and pacific oceans influencing spawning and recruitment of anguillid eels. am. fish. soc. symp. 69:231-249. miller mj, westerberg h, sparholt h, wysujack k, sørensen sr, marohn l, jacobsen mw, freese m, ayala dj, pohlmann jd, svendsen jc, watanabe s, andersen l, møller pr, tsukamoto k, munk p, hane r, 2019. spawning by the european eel across 2000 km of the sargasso sea. biol. lett. 15:20180835. moccia d, salvadori l, ferrari s, carucci a, pusceddu a, 2020. implementation of the eu ecological flow policy in italy with a focus on sardinia. adv. oceanogr. limnol. 2020:11 montgomery dr, 2003. king of fish: the thousand-year run of salmon. westview press. boulder, colo., usa: 304 pp. moore ka, jarvis jc, 2008. environmental factors affecting recent summertime eelgrass diebacks in the lower chesapeake bay: implications for long-term persistence. j. coast. res. spl. issue 55:135e147. morais p, daverat f, 2016. an introduction to fish migration. boca raton, florida (usa): crc press: 315 pp. moriarty c, dekker w, 1997. management of the european eel. fish. bull. 15:1-110. moriarty c, mccarthy d, 1982. eel, p.3-6. in: eifac technical paper no. 42 european inland fisheries advisory commission (ed.), report of the symposium on stock enhancement in the management of freshwater fisheries. mueller r, liss s, deng zd, 2019. implantation of a new micro acoustic tag in juvenile pacific lamprey and american eel. j. vis. exp. 145:e59274. myers ga, 1949. usage of anadromous, catadromous and allied terms for migratory fishes. copeia 1949:89-97. naisbett-jones lc, putman nf, stephenson jf, ladak s, young ka, 2017. a magnetic map leads juvenile european eels to the gulf stream. curr. biol. 27:1236-1240. naismith ia, knights b, 1990. studies of sampling methods and of techniques for estimating populations of eels, anguilla anguilla l. aquac. fish. manag. 21:357–367. nunn, a.d., cowx, i.g., 2012. restoring river connectivity: prioritizing passage improvements for diadromous fishes and lampreys. ambio 41:402–409. noaa. 2017. fisheries economics of the united states: 169 pp. https://www.fisheries.noaa.gov/feature-story/fisheries-economicsunited-states-2015. oliveira k, 1997. movements and growth rates of yellow-phase american eels in the annaquatucket river, rhode island. trans. amer. fish. soc. 126:638-646. ospar. 2010. background document for european eel anguilla anguilla. ospar commission: 29 pp. ovidio m, seredynski al, philippart jc, matondo bn, 2013. a bit of quiet between the migrations: the resting life of the european eel during their freshwater growth phase in a small stream. aquat. ecol. 47:291-301. ovidio m, tarrago-bès f, matondo bn, 2015. short-term responses of glass eels transported from uk to small belgian streams. ann. limnol. int. j. limnol. 51:219-226. pacariz s, westerberg h, björk g, 2014. climate change and no nco mm er cia l u se on ly c. podda et al.62 passive transport of european eel larvae. ecol. freshw. fish 23:86–94. padgett te, thomas re, borman dj, mould dc, 2020. individual-based model of juvenile eel movement parametrized with computational fluid dynamics-derived flow fields informs improved fish pass design. r. soc. open sci. 7:191505. palm s, dannewitz j, prestegaard t, wickstrom h, 2009. panmixia in european eel revisited: no genetic difference between maturing adults from southern and northern europe. heredity 103:82-89. pedersen mi, 1998. recapture rate, growth and sex of stocked cultured eels anguilla anguilla (l.). bull. français pêche piscic. 349:153-162. pedersen mi, 2000. long-term survival and growth of stocked eel, anguilla anguilla (l.), in a small eutrophic lake. dana 12:71-76. pedersen mi, 2009. does stocking of danish lowland streams with elvers increase european eel populations? am. fish. soc. symp. 58:149-156. pedersen mi, rasmussen gh, 2016. yield per recruit from stocking two different sizes of eel (anguilla anguilla) in the brackish roskilde fjord. ices j. mar. sci. 73:158-164. perkin js, gido kb, 2012. fragmentation alters stream fish community structure in dendritic ecological networks. ecol. appl. 22:2176-2187. pike c, crook v, gollock m, 2020. anguilla anguilla. in: the iucn red list of threatened species: e.t60344a 152845178. piper at, svendsen jc, wright rm, kemp ps, 2017. movement patterns of seaward migrating european eel (anguilla anguilla) at a complex of riverine barriers: implications for conservation. ecol. freshwater fish 26:87-98. piper at, whitec pr, wrightd rm, leightonc tg, kempa ps, 2019. response of seaward-migrating european eel (anguilla anguilla) to an infrasound deterrent. ecol. eng. 127:480-486. piper at, wright rm, walker am, kempps, 2013. escapement, route choice, barrier passage and entrainment of seaward migrating european eel, anguilla anguilla, within a highly regulated lowland river. ecol. eng. 57:88-96. podda c, palmas f, frau g, chessa g, culurgioni j, diciotti r, fois n, sabatini a, 2020. environmental influences on the recruitment dynamics of juvenile european eels, anguilla anguilla, in a small estuary of the tyrrhenian sea, sardinia, italy. aquatic conserv.: mar. freshw. ecosyst. 30:1638-1648. podgorniak t, blanchet s, de oliveira e, daverat f, pierron f, 2016. to boldly climb: behavioural and cognitive differences in migrating european glass eels. r. soc. open sci. 3:150665. polyakov iv, bhatt us, simmons hl, walsh d, walsh je, zhang x, 2005. multidecadal variability of north atlantic temperature and salinity during the twentieth century. j. clim. 18:4562-4581. prigge e, marohn l, hanel r, 2013. tracking the migratory success of stocked european eels anguilla anguilla in the baltic sea. j. of fish biol. 82:686-699. prouzet p, 2002. historique des captures de civelles, intensité actuelle de leur exploitation, variation de leur capturabilité par la pêche professionnelle maritime et indices de colonisation sur le bassin versant de l’adour. technical report, ifremer. http://www.ifremer.fr/indicang/boite-bassins-versants/pdf/historique-capture-civelle.pdf. psuty i, draganik b, 2008. the effectiveness of glass eel stocking in the vistula lagoon, poland. aiep 2:103-111. radinger j, hölker f, horky p, slavìk o, dendoncker n, wolter c, 2016. synergistic and antagonistic interactions of future land use and climate change on river fish assemblages. glob. change biol. 22:1505-1522. raeymaekers ja, maes ge, geldof s, hontis i, nackaerts k, volckaert fa, 2008. modeling genetic connectivity in sticklebacks as a guideline for river restoration. evol. appl. 1:475-488. réveillac é, feunteun e, berrebi p, gagnaire pa, lecomtefiniger r, bosc p, robinet t, 2008. anguilla marmorata larval migration plasticity as revealed by otolith microstructural analysis. can. j. fish. aquat. sci. 65:2127-2137. réveillac é, robinet t, rabenevanana mw, valade p, feunteun e, 2009. clues to the location of the spawning area and larval migration characteristics of anguilla mossambica as inferred from otolith microstructural analyses. j. fish. biol. 74:1866-1877. righton d, westerberg h, feunteun e, okland f, gargan p, amilhat e, metcalfe j, lobon-cervia j, sjo berg n, simon j, acou a, vedor m, walker a, trancart t, brämick u, aarestrup k, 2016. empirical observations of the spawning migration of european eels: the long and dangerous road to the sargasso sea. sci. adv. 2:e1501694. ringuet s, muto f, raymakers c, 2002. eels: their harvest and trade in europe and asia. traffic bulletin 19:2-27. robinet t, lecomte-finiger r, escoubeyrou k, feunteun e, 2003. tropical eels anguilla spp. recruiting to réunion island in the indian ocean: taxonomy, patterns of recruitment and early life histories. mar. ecol. prog. ser. 259:263-272. robinet t, réveillac e, kuroki m, aoyama j, tsukamoto k, rabenevanana mw, valade p, gagnaire pa, berrebi p, feunteun e, 2008. new clues for freshwater eels (anguilla spp.) migration routes to eastern madagascar and surroundin islands. mar. biol. 154:453-463. rochard e, pellegrini p, marchal j, béguer m, ombredane d, lassalle g, menvielle e, bagliniére jl, 2009. identification of diadromous fish species on which to focus river restoration: an example using an eco-anthropological approach (the seine basin, france). challenges for diadromous fishes in a dynamic global environment. am. fish. soc. symp. 69:691-711. rosell r, evans d, allen m, 2005. the eel fishery in lough neagh, northern ireland an example of sustainable management? fish. manag. ecol. 12:377-385. rossier o, 1997. comparison of gillnet sampling and night visual census of fish communities in the littoral zone of lake geneva, switzerland. arch. hydrobiol. 139:223-233. sanderson fj, donald pf, pain dj, burfield ij, van bommel fpj, 2006. long-term population declines in afro-palearctic migrant birds. biol. conserv. 131:93-105. sandlund ot, diserud poole oh, bergesen k, dillane m, rogan g, durif c, thorstad eb, vøllestad la, 2017. timing and pattern of annual silver eel migration in two european watersheds are determined by similar cues. ecol. evol. 7:5956-5966. schabetsberger r, økland f, aarestrup k, kalfatak d, no nco mm er cia l u se on ly hard times for the catadromous european eel 63 sichrowsky u, tambets m, dall’oimo g, kaiser r, miller p, 2013. oceanic migration behaviour of tropical pacific eels from vanuatu. mar. ecol. prog. ser. 475:177-190. schabetsberger r, okland f, kalfatak d, sichrowsky u, meelis t, aarestrup k, gubili c, sarginson j, boufana b, jehle r, dall’olmo g, miller mj, scheck a, kaiser r, quartly g, 2015. genetic and migratory evidence for sympatric spawning of tropical pacific eels from vanuatu. mar. ecol. prog. ser. 521:171-187. schabetsberger r, scheck a, kaiser r, leaana r, gubili c, økland f, 2019. oceanic migration behaviour of pacific eels from samoa. fish. manag. ecol. 26:53-56. schiavina m, bevacqua d, melià p, crivelli aj, gatto m, de leo ga, 2015. a user-friendly tool to assess management plans for european eel fishery and conservation. environ. model. softw. 64:9-17. schmidt j, 1922. the breeding places of the eel. philos. trans. r. soc. 211:179-208. schmucker ak, johnson ns, galbraith hs, li w, 2016. glasseel-stage american eels respond to conspecific odor as a function of concentration. t. am. fish. soc. 145:712-722. schulze t, kahl u, radke rj, benndorf j, 2004. consumption, abundance and habitat use of anguilla anguilla in a mesotrophic reservoir. j. fish biol. 65:1543-1562. secor dh, 2015. migration ecology of fishes. baltimore, usa: john hopkins university press: 304 pp. shiao jc, ložys l, iizuka y and tzeng wn, 2006. migratory patterns and contribution of stocking to the population of european eel in lithuanian waters as indicated by otolith sr:ca ratios. j. fish biol. 69:749-769. shields fd, knight ss, cooper m, 1995. incised stream physical habitat restoation with stone weirs. regul. rivers res. manag. 10:181-198. sjöberg nb, wickström h, asp a, petersson e, 2017. migration of eels tagged in the baltic sea and lake mälaren in the context of the stocking question. ecol. freshw. fish 26:517-532. simon j, dörner h, 2014. survival and growth of european eels stocked as glassand farm-sourced eels in five lakes in the first years after stocking ecol. freshw. fish 23:40-48. simon j, dörner h, scott rd, schreckenbach k, knösche r, 2013. comparison of growth and condition of european eels stocked as glass and farm sourced eels in lakes in the first four years after stocking. j. appl. ichthyol. 29:323-330. smith km, byron cj, sulikowski ja, 2016. modeling predator-prey linkages of diadromous fishes in an estuarine food web. mar. coast. fish. 8:476-491. sola c, 1995. chemoattraction of upstream migrating glass eels anguilla anguilla to earthy and green odorants. env. biol. fishes 43:179-185. sola c, tongiorgi p, 1996. the effects of salinity on the chemotaxis of glass eels, anguilla anguilla, to organic earthy and green odorants. env. biol. fishes 47:213-218. sorensen pw, 1986. origins of the freshwater attractant(s) of migrating elvers of the american eel, anguilla rostrata. env. biol. fishes 17:185-200. stacey ja, pratt tc, verreault g, fox mg, 2015. a caution for conservation stocking as an approach for recovering atlantic eels. aquat. conserv. mar. freshw. ecosyst. 25:569-580. taberlet p, coissac e, pompanon f, brochmann c, willerslev e, 2012. towards next-generation biodiversity assessment using dna metabarcoding. mol. ecol. 21:2045-2050. teichert n, tetard s, trancart t, feunteun e, acou a, de oliveira e, 2020. resolving the trade-off between silver eel escapement and hydropower generation with simple decision rules for turbine shutdown. j. env. manag. 261:110212. tesch fw, 1977. the eel. biology and management of anguillid eels. chapman & hall, london: 370 pp. tesch fw, 2003. the eel. blackwell science: 408 pp. thorstad eb, rikardsen ah, alp a, økland f, 2013. the use of electronic tags in fish research – an overview of fish telemetry methods. turkish j. fish. aquat. sci. 13:881-896. tosi l, sola c, 1993. role of geosmin, a typical inland water odour, in guiding glass eel anguilla anguilla (l.) migration. ethology 95:177-185. tosi l, spampanato a, sola c, tongiorgi p, 1989. relation of water odour, salinity and temperature to ascent of glass-eels, anguilla anguilla (l.): a laboratory study. j. fish biol. 36:327-340. trancart t, acou a, de oliveira e, feunteun e, 2013. forecasting animal migration using sarimax: an efficient means of reducing silver eel mortality caused by turbines. endanger. species res. 21:181-190. trancart t, carpentier a, acou a, charrier f, mazel v, danet v, feunteun e, 2020. when “safe” dams kill: analyzing combination of impacts of overflow dams on the migration of silver eels. ecol. eng. 145:105741. trancart t, lambert p, daverat f, rochard e, 2014. from selective tidal transport to counter-current swimming during watershed colonisation: an impossible step for young-of-theyear catadromous fish? knowl. manag. aquat. ecosyst. 2014:412. trancart t, tétard s, acou a, feunteun e, schaeffer f, de oliveira e, 2018. silver eel downstream migration in the river rhine, route choice, and its impacts on escapement: a 6-year telemetry study in a highly anthropized system. ecol. eng. 123:202-211. tsukamoto k, 1990. recruitment mechanism of the eel, anguilla japonica, to the japanese coast. j. fish. biol. 36:659-671. tsukamoto k, aoyama j, 1998. evolution of the freshwater eels of the genus anguilla: a probable scenario environ. biol. fish. 52:139-148. van ginneken vjt, maes ge, 2005. the european eel (anguilla anguilla, linnaeus), its lifecycle, evolution and reproduction: a literature review. rev. fish. biol. fish 15:367-398. van puijenbroek pjtm, buijse ad, kraak mhs, verdonschot pfm, 2019. species and river specific effects of river fragmentation on european anadromous fish species. river res. appl. 35:68-77. verreault g, mingelbier m, dumont p, 2012. spawning migration of american eel anguilla rostrata from pristine (1843– 1872) to contemporary (1963–1990) periods in the st lawrence estuary, canada. j. fish. biol. 81:387-407. vörösmarty cj, mcintyre pb, gessner mo, dudgeon d, prusevich a, green p, glidden s, bunn se, sullivan ca, reidy liermann c davies pm, 2010. global threats to human water security and river biodiversity. nature 467:555-561. walker am, andonegi e, apostolaki p, aprahamian m, beaulaton l, bevacqua d, bevacqua p, briand c, cannas a, de eyto e, dekker w, de leo g, diaz e, doering-arjes p, fladung e, jouanin c, lambert p, poole r, oeberst r, schino nco mm er cia l u se on ly c. podda et al.64 avina m, 2013. lot 2: pilot project to estimate potential and actual escapement of silver eel. final project report, service contract s12.539598, studies and pilot projects for carrying out the common fisheries policy. brussels, european commission, directorate general for maritime affairs and fisheries (dg mare): 358 pp. wang ch, tzeng wn, 2000. the timing of metamorphosis and growth rates of american and european eel leptocephali: a mechanism of larval segregative migration. fish. res. 46:191-205. weber m, 1986. fishing method and seasonal occurrence of glass eels (anguilla anguilla, l.) in the rio minho, west coast of the iberian peninsula. vie milieu 366:243-250. westerberg h, miller mj, wysujack k, marohn l, freese m, pohlmann jd, watanabe s, tsukaoto k, hanel r, 2018. larval abundance across the european eel spawning area: an analysis of recent and historic data. fish fish. 19:890-902. westerberg h, sjöberg n, lagenfelt i, aarestrup k, righton d, 2014. behaviour of stocked and naturally recruited european eels during migration. mar. ecol. progr. ser. 496:145-157. westerberg ik, wagener t, coxon g, mcmillan hk, castellarin a, montanari a, freer j, 2016. uncertainty in hydrological signatures for gauged and ungauged catchments, water resour. res. 52:1847-1865. westin l, 1998. the spawning migration of european silver eel (anguilla anguilla l.) with particular reference to stocked eel in the baltic. fish. res. 38:257-270. westin l, 2003. migration failure in stocked eels anguilla anguilla. mar. ecol. pr. ser. 254:307-311. wickström h, sjöberg nh, 2014. traceability of stocked eels— the swedish approach. ecol. freshw. fish 23:33-39. wickström h, westin l, clevestam p, 1996. the biological and economic yield from a long-term eel-stocking experiment. ecol. freshw. fish 5:140-147. wilcove ds, wikelski m, 2008. going, going, gone: is animal migration disappearing. plos biol. 6:e188. wilkes ma, mckenzie m, webb ja, 2018. fish passage design forsustainable hydropower in the temperate southern hemisphere: an evidence review. rev. fish biol. fish. 28:117-135. winter h, jansen h, bruijs m, 2006. assessing the impact of hydropower and fisheries on downstream migrating silver eel, anguilla anguilla, by telemetry in the river meuse. ecol. freshw. fish 15:221-228. wofford je, gresswell re, banks ma, 2005. influence of barriers to movement on within�watershed genetic variation of coastal cutthroat trout. ecol. appl. 15:628-637. wolter c, 2015. historic catches, abundance, and decline of atlantic salmon salmo salar in the river elbe. aquat. sci. 77:367-380. wysujack k, westerberg h, aarestrup k, trautner j, kurwie t, nagel f, hanel r, 2015. the migration behaviour of european silver eels (anguilla anguilla) released in open ocean conditions. mar. freshw. res. 66:145-157. zarfl c, lumsdon ae, berlekamp j, tydecks l, tockner k, 2015. a global boom in hydropower dam construction. aquat. sci. 77:161-170. zompola s, katselis g, koutsikopoulos c, cladas y, 2008. temporal patterns of glass eel migration (anguilla anguilla l. 1758) in relation to environmental factors in the western greek inland waters. estuar. coast. shelf sci. 80:330-338. no nco mm er cia l u se on ly layout 1 introduction italian inland waters are known to be heavily affected by the occurrence of non-indigenous species (nis) (gherardi et al., 2007). these biological invaders are one of the major threats to native freshwater fauna, altering the habitat structure and assemblage composition and ultimately leading to a significant loss of native biodiversity (naselliflores and marrone, 2019 and references therein). italy is characterized by a rich fauna of cladocera, which includes 17 daphnia o.f. müller, 1785 species belonging to the subgenera daphnia s.s. and ctenodaphnia, plus several subspecies or hybrid taxa of dubious taxonomical value (ruffo and stoch, 2005; marrone et al., 2007). within the species belonging to the genus daphnia s.s. occurring in italy, two are allochthonous taxa of nearctic origin, i.e., d. ambigua scourfield, 1947 and d. parvula fordyce, 1901 (margaritora, 1985; riccardi et al., 2004). furthermore, the presence of a non-native lineage of american origin within d. pulex leydig, 1860 was recorded in sardinia and piedmont (fadda et al., 2011; marková et al., 2013). to date, the only non-native cladocerans positively known to occur in sicily are daphnia parvula and d. ambigua (marrone et al., 2006; marrone and naselli-flores, 2015), although also d. galeata g.o. sars, 1864 and d. cucullata g.o. sars, 1862 have a dubious status in the region (marrone and vecchioni, 2020). moreover, no molecular data are to date available on the sicilian populations of d. pulex, so that no information about the native or non-native status of sicilian d. pulex is available. furthermore, it must be taken into account that the taxonomy of the daphnia pulex complex is rather difficult due to the absence of clear morphological features distinguishing among taxa listed under daphnia pulex (condeporcuna et al., 2020 and reference therein), a binomen which is misleadingly used for distinct different evolutionary lineages belonging to different biogeographical regions (mergeay et al., 2008; crease et al., 2012; ma et al., 2019). among them, the nearctic lineage has been introduced in several countries, attaining an almost worldwide distribution with the sole exception of antarctica (crease et al., 2012; conde-porcuna et al., 2020). in the light of the evidence of the presence of the allochthonous north american lineage of daphnia pulex in italian mainland and sardinia (fadda et al., 2011; marková et al., 2013, 2017), the aim of this work was to identify the daphnia lineages occurring in the sicilian populations using a mitochondrial molecular marker, and thus verifying whether the evolutionary lineages occurring in sicily belong to the allochthonous nearctic or the autochthonous european groups. accordingly, we carried out samplings in a selection of the known sites of occurrence of the species, and in an unpublished site where the species was found to occur. methods based on the data available in the literature, daphnia pulex is currently known for 19 sites in sicily, which are limited to medium-high altitudes of the nebrodi area, madonie area, etna and bosco della ficuzza (marrone and vecchioni, 2020). the populations from which samples of daphnia pulex to be molecularly characterized article molecular data attest to the occurrence of autochthonous daphnia pulex (crustacea, branchiopoda) populations in sicily, italy luca vecchioni,* marco arculeo, federico marrone department of biological, chemical and pharmaceutical sciences and technologies, university of palermo, via archirafi 18, 90123 palermo, italy abstract biological invasions are known to be among the most important threats to the long-term conservation of native biota, and their effects might be even more difficult to contrast when they are cryptic, i.e., when the non-native invaders cannot be easily recognised based on morphology, and can thus be confused with native taxa. such cryptic invasions are known to widely occur in the cladoceran genus daphnia o.f. müller, 1785, so that the actual distribution and status of most species and lineages need to be checked with a genetic approach. in the frame of this work, we investigated if the sicilian populations of d. (daphnia) pulex leydig, 1860 belonged to the allochthonous north american lineage, which is known to occur in several regions of the palearctic and afrotropical biogeographical regions, or rather to the autochthonous european lineage of the species. the molecular results obtained, based on a fragment of the mitochondrial gene encoding for nadh subunit dehydrogenase 5 (nd5), allowed us to rule out the allochthonous status of the species, confirming the presence of autochthonous relictual lineages of d. pulex in sicily. the native status of these populations is in agreement with their local distribution, limited to natural and poorly-impacted water bodies mostly located in wooded areas at medium and high altitudes. the current local distribution of d. pulex in sicily is possibly linked to the end of the last glacial maximum and the onset of warmer climatic conditions in the early holocene, which led the species to take refuge in colder microthermal refugia located at high altitudes, determining their current relictual distribution. no nco mm er cia l u se on ly l. vecchioni et al.30 were collected were chosen based on the known occurrence sites, with the aim of including in the analyses at least a population from each of the four major distribution subareas of the species on the island. in addition, one unpublished site of occurrence was found in the context of this sampling campaign (“me076” – 37.94158 n, 14.683676 e – 1559 m a.s.l.). a map of the known and sampled occurrence sicilian sites of the species was made using qgis software v. 3.18 (http://www.qgis.org). microcrustaceans were collected both in open waters and in the littoral areas of each water body through the use of handand\or towing-plankton nets with a mesh size of 125-200 μm, depending on size, depth and vegetation cover of the sampled sites. the collected samples were fixed in situ with 96% ethanol and microcrustaceans were sorted out in laboratory under a stereomicroscope. cladocera were morphologically identified according to margaritora (1985), alonso (1996), and benzie (2005). all the studied samples are currently stored in fm’s branchiopod collection at the university of palermo, italy, and are available for loan on request. a single d. pulex individual from each population was carefully cleaned of any impurities and soaked in doubledistilled water for 5-10 minutes in order to eliminate the residual ethanol, and then processed for dna extraction using the bioron gmbh “ron’s tissue dna mini kit”, following the manufacturer’s instructions. the selective amplification of a fragment from the gene encoding nadh subunit dehydrogenase 5 (nd5), was carried out by polymerase chain reaction (pcr) using the primers nd5newf (5’-aaa cct cta aab ttc yka rct3 ‘) and nd5newr (5’-cat rtt yat rtc rgg ggt tgt3’), described by dufresne et al. (2011). the pcr mix consisted of 18.9 μl of distilled water, 2.5 μl of buffer 10x which includes 15 mm of mgcl2, 0.4 μl of dntps (10 mm for each), 0.4 μl of each of the primers (10 μm), 0.4 μl of taq polymerase (5 u / μl) and 2 μl of template dna, for a total volume of 25 μl. the thermal cycle consisted of 35 cycles of denaturation (94°c for 1 min), annealing (50°c for 1 min) and extension (72°c for 1 min), followed by five minutes at 72°c for the final extension step. after pcr, 5 μl of each pcr product were used to perform electrophoresis on 2% agarose gel at 90 v for 20 min and then visualized with a uv transilluminator. when pcr products showed a clear single band, of the expected length, they were purified using the exo-sap-it® kit (affymetrix usb, santa clara, ca, usa). sequencing was performed by macrogen inc. (madrid, spain; https://dna.macrogen.com/eng/) using an abi 3130xl (applied biosystems, waltham, ma, usa) sequencer. the same primers used previously for pcr were subsequently used for direct sequencing of the pcr products. the quality of the obtained chromatograms was checked through the measurement of their “phred score” (richterich, 1998). only those sequences that showed continuous high quality base readings (qv > 20) were used. chromatograms were analysed and manually proofread using the software chromas software v. 2.6.2 (technelysium, pty. ltd., south brisbane, australia). overall, six novel nd5 sequences of daphnia pulex were produced. moreover, in order to compare the new sequences with those publicly available, ten daphnia pulex sequences, eight d. pulicaria forbes, 1893 sequences, four d. tenebrosa sars, 1898 sequences and one d. magna straus, 1820 sequence (used as an outgroup) were downloaded from genbank and included in the analyses (see tab. 1 for their accession number, a.n.). all sequences were aligned with megax software (kumar et al., 2018) using the clustalw method (thompson et al., 1997). mrbayes software v. 3.2.6 (ronquist et al., 2012) and phyml v. 3 (guindon and gascuel, 2003) were used for inferring the molecular identification and phylogenetic relationships between taxa, using bayesian inference (bi) and maximum likelihood (ml) analyses. as support measures for the nodes, bootstrap values (felsenstein, 1985) were calculated with 1000 replicates in ml trees, while the posterior probability values were reported in the bi tree. partitionfinder v. 1.0.1 (lanfear et al., 2012) was used to choose the best evolutionary model following the “akaike information criterion” (aic; akaike, 1974). a hasegawa – kishino – yano evolutionary sequence model with a proportion of gamma and invariant sites (hky+i+γ; nst = 2) was used in the bi and ml analyses. in the bi analysis, two independent markov chain monte carlo analyses were performed with 1 million generations (temp.: 0.2; default priors). the trees and parameter values were sampled every 100 generations, resulting in 10,000 trees for each analysis. the convergence in the analysis was reached (effective sample size (ess) greater than 533.13 in all the analyses performed). the initial 25% of trees were discarded as “burn-in”. results in addition to five of the published sites from the madonie, nebrodi and bosco della ficuzza areas, the species was collected in an novel site on the nebrodi mountains (me076), located close to the site “me004”, where the species was already known to occur (see tab. 1 in marrone and vecchioni, 2020). in all the sampled populations, parthenogenetic females coexisted with males and ephippial females. unfortunately, logistic constraints made not possible to collect fresh samples in the only known site for the species within the etna area (ct011, see tab. 1 in marrone and vecchioni, 2020), which was thus not included in the analyses. no nco mm er cia l u se on ly molecular characterization of daphnia pulex in sicily 31 overall, six new sequences belonging to daphnia pulex were produced from six different sicilian water bodies (fig. 1, tab. 1). after having trimmed out the sequences, a properly aligned fragment of 624 bp long of the nd5 mtdna gene was obtained. novel sequences were deposited in genbank (a.n., mz489122mz489127). the bi and ml trees based on the mitochondrial nd5 mtdna fragment and rooted on d. magna showed a congruent and well supported topology highlighting how the analysed sequences of daphnia pulex s.l. create two paraphyletic clades: the first including north american d. pulex populations and the allochthonous north american d. pulex populations occurring in europe (“napx”); the second one includes the autochthonous european d. pulex populations (“epx”) including the analysed sicilian populations (i.e., pa074, pa079, pa081, me004, me013, me076; tab. 1). all the sicilian sequences clustered together in a monophyletic clade showing only two slightly different haplotypes (i.e., there is only a single base of difference between the sequence that belongs to pa079 and the remaining ones). similarly to what was observed in d. pulex, also d. pulicaria shows a marked paraphyly (fig. 2) that separates the north american populations (“napc”) from the european ones (“epc”). discussion based on the molecular evidence here reported, the sicilian populations of daphnia pulex unequivocally belong to the native european lineage. taking into account the distribution of the species in sicily, this result is not surprising. in fact, sicilian d. pulex populations seem to be linked to natural, poorlytab. 1. origin and genbank accession numbers (a.n.) for the analysed daphnia specimens. geographic coordinates are expressed as decimal degrees (map datum: wgs84). taxon country location latitude (n) longitude (e) a.n. clade reference d. pulicaria albania ohrid 40.95 20.71 kc536551 epc marková et al., 2013 switzerland alps 46.67 8.04 kc536536 epc marková et al., 2013 czech republic bohdaneč 49.78 15.23 kc536543 epc marková et al., 2013 italy dolomites 46.48 11.66 kc536525 epc marková et al., 2013 greenland nuuk 64.15 -51.31 kc536586 epc marková et al., 2013 canada winnipeg, manitoba 52.12 -97.25 kc536611 napc marková et al., 2013 usa indiana 39.9 -85.43 kc536600 napc marková et al., 2013 iceland reykjavik 64.13 -21.94 kc536621 napc marková et al., 2013 d. pulex czech republic blatná 49.42 13.78 kc536544 epx marková et al., 2013 germany regensburg 49 12.15 kc536593 epx marková et al., 2013 uk streetly end 52.14 0.35 kc536555 epx marková et al., 2013 lithuania vilnius 54.75 25.29 kc536560 epx marková et al., 2013 canada listowel pond, ontario 43.73 -80.95 kc536602 napx marková et al., 2013 canada disputed road, ontario 42.22 -83.03 hq434640 napx vergilino et al., 2011 canada res. duchesnier, quebec 48.13 -68.63 kc536603 napx marková et al., 2013 italy pa079 37.889944 13.394575 mz489122 epx present work italy pa074 37.823410 14.127241 mz489123 epx present work italy pa081 37.901131 13.408438 mz489124 epx present work italy me013 37.951944 14.698331 mz489125 epx present work italy me076 37.941580 14.683670 mz489126 epx present work italy me004 37.939456 14.682634 mz489127 epx present work italy avigliana 45.070000 7.390000 kc536565a napx marková et al., 2013 italy sardinia 40.560000 9.320000 kc536565b napx marková et al., 2013 italy bodrio del pastore iii 45.001389 10.323889 kr233296 epx marková et al., 2017 spain lake borreguil, sierra nevada 37.052 -03.30 mw883468 napx conde-porcuna et al., 2021 sweden hq434644 epx vergilino et al., 2011 kenya lake naivasha -0.771667 36.361667 dq235240 napx mergeay et al., 2006 south africa cape flats -33.98 18.656667 dq235233 napx mergeay et al., 2006 kenya lake baringo 0.535764 36.063759 dq235242 napx mergeay et al., 2006 japan kosugi, toyama 36.68 137.09 jx532913 napx crease et al., 2012 japan higashi hiroshima, hiroshima 34.35 132.8 jx532907 napx crease et al., 2012 china lake wusutu 40.86 111.55 mh632088 napx ma et al., 2019 d. tenebrosa canada churchill, manitoba 58.77 -94.17 kc536605 marková et al., 2013 russia petchora delta 68.06 53.58 kc536564 marková et al., 2013 svalbard storvatnet, ny-alesund 78.92 11.88 kc536580 marková et al., 2013 russia tsvetkov cape, taimyr 74.92 112.62 kc536608 marková et al., 2013 d. magna mt199637 lee (unpublished) no nco mm er cia l u se on ly l. vecchioni et al.32 mineralized ponds located within wooded areas at medium and high altitudes (marrone and vecchioni, 2020), where phenomena of biological invasions are less frequent than in disturbed water bodies located in highlyanthropized areas. in fact, in italy, the non-native d. pulex lineage was to date found in artificial or highly-anthropized water bodies both in sardinia and piedmont (fadda et al., 2011; marková et al., 2017). in light of these results, the hypothesis by marrone et al. (2009) that sicilian d. pulex populations are relictual elements that colonised sicily during the late pleistocene glacial events, coming from the balkan peninsula or northern italy, and which later found a “cool” refuge at higher altitude with the onset of warmer and low-moisture climatic conditions in the early holocene (curry et al., 2016), seems to be supported. interestingly, based on the currently available data, it seems that native european d. pulex populations preferentially inhabit small, natural water bodies, whereas the populations of the non-native american lineage are mostly occurring in artificial or highly-disturbed habitats (schwenk et al., 2000; fadda et al., 2011; vergilino et al., 2011) at medium and low altitude with a single high-altitude exception reported for a spanish lake located in the sierra nevada (conde-porcuna et al., 2020). native and non-native d. pulex lineages are known to coexist sympatrically, but not syntopically, in the po river basin, with the native populations occurring in the “bodri” (i.e., temporary natural ponds) (marková et al., 2017), and the nonnative one found in the highly-anthropized avigliana lakes (virgilino et al., 2011; marková et al., 2013). conversely, for sardinian populations molecular data are available only from a single site, i.e., an artificial reservoir built in sos canales (fadda et al., 2011). however, d. pulex is also occurring in smaller, natural astatic and temporary ponds located in the southern part of the island (margaritora et al., 2021; marrone and stoch, unpublished data), and it should be checked whether the d. pulex populations inhabiting these marginal, natural habitats belong to the alien or the native lineage. fig. 1. geographic location of daphnia pulex sampling sites for which molecular data are available. red circles indicate sites where the north american lineages, “napx”, of daphnia pulex occur. green circles and diamonds indicate where the european lineages, “epx”, of the species occur. green diamonds indicate the novel sampled sites. due to the scale of the map some of the novel sites overlap and thus are not displayed. see tab. 1 for the coordinates of the sampling sites and for more information on the collected species. no nco mm er cia l u se on ly molecular characterization of daphnia pulex in sicily 33 fig. 2. bayesian phylogram of daphnia spp. based on the 624 bp fragment of the mtdna nd5. d. magna was used as outgroup. node statistical support is reported as nodal posterior probabilities (bayesian inference of phylogeny, bi)/bootstrap values (maximum likelihood, ml). *, nodal statistical supports <0.50. square brackets group the samples according to the current taxonomy of the genus. novel sequences are reported in bold. napx, north american pulex; napc, north american pulicaria; epx, european pulex; epc, european pulicaria. the analysed specimens are reported using the codes listed in tab. 1. no nco mm er cia l u se on ly l. vecchioni et al.34 conclusions this study provides the first molecular data about the sicilian populations of the water flea daphnia pulex. all the sampled populations proved to belong to the autochthonous european lineage of the species, in accordance with their previously hypothesised native and relictual status (marrone et al., 2009; marrone and vecchioni, 2020). sicilian d. pulex populations are thus important management units, to be attentively managed and preserved. considering the threats to native taxa linked with the occurrence and spreading of invasive, alien species (marrone and naselli-flores, 2015), and the importance of timely acting to have a chance to effectively manage the biological invasions, the molecular characterization of the d. pulex populations occurring in southern sardinia, peninsular italy, and malta (margaritora et al., 2021) is urgent and desirable. acknowledgments sara la franca is acknowledged for the help she provided during the preparation of this manuscript. two anonymous reviewers are acknowledged for their comments to a first draft of the manuscript. conflict of interest the authors declare no competing interests. corresponding author: luca.vecchioni@unipa.it keywords: cladocera; anomopoda; biological invasions; cryptic invasions. received: 1 july 2021. accepted: 14 september 2021. this work is licensed under a creative commons attribution noncommercial 4.0 license (cc by-nc 4.0). ©copyright: the author(s), 2021 licensee pagepress, italy advances in oceanography and limnology, 2021; 12:9947 doi: 10.4081/aiol.2021.9947 references akaike h, 1974. a new look at the statistical model identification. ieee trans. automat. contr. 19:716–723. alonso m, 1996. crustacea, branchiopoda. fauna iberica 7. museo nacional de ciencias naturales, csic, madrid, 486 pp. benzie jah, 2005. cladocera: the genus daphnia (including daphniopsis). guides to the identification of the microinvertebrates of the continental waters of the world. backhuys publishers, leiden: 376 pp. conde-porcuna jm, veiga j, moreno e, jiménez l, ramos-rodríguez e, pérez-martínez c, 2021. spatiotemporal genetic structure in the daphnia pulex complex from sierra nevada lakes (spain): reproductive mode and first record of north american d. cf. pulex in european alpine lakes. j. plankton res. 43:380–395. crease tj, omilian ar, costanzo ks, taylor dj, 2012. transcontinental phylogeography of the daphnia pulex species complex. plos one 7:e46620. curry b, henne pd, mesquita-joanes f, marrone f, pieri v, la mantia t, calò c, tinner w, 2016. holocene paleoclimate inferred from salinity histories of adjacent lakes in southwestern sicily (italy). quat. sci. rev. 150: 67-83. dufresne f, marková s, vergilino r, ventura m, kotlík p, 2011. diversity in the reproductive modes of european daphnia pulicaria deviates from the geographical parthenogenesis. plos one 6:e20049. fadda a, marková s, kotlík p, lugliè a, padedda b, buscarinu p, sechi n, manca m, 2011. first records of planktonic crustaceans in sardinian reservoirs. biologia 66:856–865. felsenstein j, 1985. confidence limits on phylogenies: an approach using the bootstrap. evolution 39:783–791. gherardi f, bertolino s, bodon m, casellato s, cianfanelli s, ferraguti m, lori e, mura g, nocita a, riccardi n, rossetti g, rota e, scalera r, zerunian s, tricarico e, 2007. animal xenodiversity in italian inland waters: distribution, modes of arrival, and pathways. biol. invasions 10:435–454. guindon s, gascuel o, 2003. a simple, fast, and accurate algorithm to estimate large phylogenies by maximum likelihood. syst. biol. 52:696–704. kumar s, stecher g, li m, knyaz c, tamura k, 2018. mega x: molecular evolutionary genetics analysis across computing platforms. mol. biol. evol. 35:1547–1549. lanfear r, calcott b, ho syw, guindon s, 2012. partitionfinder: combined selection of partitioning schemes and substitution models for phylogenetic analyses. mol. biol. evol. 29:1695–1701. ma xl, petrusek a, wolinska j, hu w, yin mb, 2019. lineage diversity and reproductive modes of the daphnia pulex group in chinese lakes and reservoirs. mol. phylogenet. evol. 130:424–433. margaritora fg, 1985. cladocera. fauna d’italia, ed. calderini, bologna, 399 pp. margaritora fg, vagaggini d, stoch f, 2021. crustacea branchiopoda cladocera. in: bologna ma, zapparoli m, oliverio m, minelli a, bonato l, cianferoni f, stoch f. (eds.), checklist of the italian fauna. version 1.0. last update: 2021-05-31. marková s, dufresne f, manca m, kotlik p, 2013. mitochondrial capture misleads about ecological speciation in the daphnia pulex complex. plos one 8:e69497. marková s, maurone c, racchetti e, bartoli m, rossi v, 2017. daphnia diversity in water bodies of the po river basin. j. limnol. 76:261-271. marrone f, alfonso g, naselli-flores l, 2007. on daphnia (ctenodaphnia) similis claus, 1877 and other interesting anomopods (crustacea, branchiopoda) from apulia (southern italy). thalassia salentina 30: 45-55. no nco mm er cia l u se on ly molecular characterization of daphnia pulex in sicily 35 marrone f, barone r, naselli-flores l, 2006. ecological characterization and cladocerans, calanoid copepods and large branchiopods of temporary ponds in a mediterranean island (sicily, southern italy). chem. ecol. 22:181–190. marrone f, castelli g, naselli-flores l, 2009. sicilian temporary ponds: an overview on the composition and affinities of their crustacean biota, p. 189-202. in: fraga i, argiumbau p. (eds.), international conference on mediterranean temporary ponds. proceedings & abstracts. consell insular de menorca. recerca, 14. maό, menorca. marrone f, naselli-flores l, 2015. a review on the animal xenodiversity in sicilian inland waters (italy). adv. oceanogr. limnol. 6:5451. marrone f, vecchioni l, 2020. the genus daphnia in sicily and malta (crustacea, branchiopoda, anomopoda). pp. 105-124 in: la mantia et al. (eds.), life on islands, 1. biodiversity in sicily and surrounding islands. studies dedicated to bruno massa. edizioni danaus, palermo. mergeay j, aguilera x, declerck s, petrusek a, huyse t, de meester l, 2008. the genetic legacy of polyploid bolivian daphnia: the tropical andes as a source for the north and south american d. pulicaria complex. mol. ecol. 17:1789–1800. mergeay j, verschuren d, de meester l, 2006. invasion of an asexual american water flea clone throughout africa and rapid displacement of a native sibling species. proc. r. soc. lond. b. biol. sci. 273:2839–2844. naselli-flores l, marrone f, 2019. different invasibility of permanent and temporary waterbodies in a semiarid mediterranean island. inland waters 9:411–421. qgis development team (2018) qgis geographic information system. open-source geospatial foundation project. available from: http://www.qgis.org/ riccardi n, giussani g, margaritora f, couchaud b, 2004. population dynamics of the pioneer population of daphnia parvula fordyce during the invasion of lake candia (northern italy). j. limnol. 63:44–52. richterich p, 1998. estimation of errors in “raw” dna sequences: a validation study. genome res. 8:251-259. ronquist f, teslenko m, van der mark p, ayres dl, darling a, höhna s, larget b, liu l, suchard ma, huelsenbeck jp, 2012. mrbayes 3.2: efficient bayesian phylogenetic inference and model choice across a large model space. syst. biol. 61:539-542. ruffo s, stoch f, 2005. [checklist e distribuzione della fauna italiana].[article in italian]. memorie del museo civico di storia naturale di verona 2 ser. sez. sci. vita 16:1–307. schwenk k, posada d, hebert pdn, 2000. molecular systematics of european hyalodaphnia: the role of contemporary hybridization in ancient species. proc. r. soc. b 267:1833– 1842. thompson jd, gibson tj, plewniak f, jeanmougin f, higgins dg, 1997. the clustal x windows interface: flexible strategies for multiple sequence alignment aided by quality analysis tools. nucleic acids res. 25:4876–4882. vergilino r, marková s, ventura m, manca m, dufresne f, 2011. reticulate evolution of the daphnia pulex complex as revealed by nuclear markers. mol. ecol. 20:1191–1207. no nco mm er cia l u se on ly layout 1 introduction the giant red shrimp aristaeomorpha foliacea (risso, 1827) is one of the key species in mediterranean deep-sea benthic communities and it is also one of the most important target species for mediterranean bottom trawling (cau et al., 2002; guillen et al., 2012; rinelli et al., 2013, palmas et al., 2017a). the species is exploited on muddy bottoms, mainly at depths of 400-800 m, using traditional trawl nets (ragonese et al., 2002; sbrana et al., 2003; sabatini et al., 2011; sala et al., 2015). in 2016, declared landings of red shrimp (a. foliacea jointly with aristeus antennatus (risso, 1816)), amounted to 5500 t across the entire mediterranean sea (stecf, 2015). nowadays, deep-water shrimps account for about 5% of the total professional fishing income in the mediterranean sea (stecf, 2019), with increasing catches in italy and spain in particular, which are the main producers in europe (eumofa, 2019). in the mediterranean sea, giant red shrimp distribution differs among western and eastern basins according to a geographical gradient (cau et al., 2002; politou et al., 2004; cartes et al., 2011a). a. foliacea is predominant in sardinia, northern and central tyrrhenian sea, strait of sicily, ionian sea, whereas it is nearly absent in the ligurian and catalan sea, balearic islands and eastern mediterranean basin (ragonese and bianchini, 1995; papaconstantinou and kapiris, 2003). several authors have reported a strong correlation between the geographical distribution of red shrimp and environmental factors, including the geomorphological characteristics of the fishing grounds (e.g. presence of canyons and underwater reliefs) (abellò et al., 2002, sabatini et al., 2007, 2011), trophic factors (cartes et al., 2008), hydrological factors (ghidalia and bourgois, 1961; sardà et al., 2004; carney, 2005; company et al., 2008; guijarro et al., 2008; massuti et al., 2008; canals et al., 2009; cartes et al., 2011a; cartes et al., 2011b) and fishing pressure (relini and orsi relini, 1987; bianchini and ragonese, 1994; blanchard, 2001; d’onghia et al., 2005). despite these studies, many aspects of the interactions among environmental variables and spatial distribution of the species remain unclear, yet. the aim of this study is to provide further information about the possible relationships between environmental factors and the spatio-temporal distribution of giant red shrimp in sardinian waters (western mediterranean), using multivariate models (gams and regression trees). methods study area the seas around sardinia represent a particularly interesting environment, for their central position in the western mediterranean basin, its geomorphologic heterogeneity and the presence of extended fishing bottoms (cau et al., 1994). these features (i.e., submarine canyons article exploring relationships between the distribution of giant red shrimp aristaeomorpha foliacea (risso, 1827) and environmental factors in the central-western mediterranean sea cinzia podda,*§ francesco palmas,§ serenella cabiddu, paola pesci, andrea sabatini department of life and environmental sciences, university of cagliari, via fiorelli 1, 09126 cagliari, italy §cinzia podda and francesco palmas contributed equally to this work. abstract mediterranean giant red shrimp aristaeomorpha foliacea (risso, 1827) is one of the dominant species in deep-sea megafaunal assemblages, plays a key role in deep-sea communities and it is considered one of the most important targets of deep-water trawl fishing. although a large number of studies have analysed the spatial distribution of epibenthic crustaceans in bathyal habitats with respect to environmental, geomorphological and hydrological factors, as well as fishing pressure, the manner in which these variables synergistically affect the spatio-temporal changes of giant red shrimp is unclear. to analyse the possible effects of abiotic predictors on the spatio-temporal distribution of giant red shrimp, generalized additived models (gams) and regression trees were produced. biological data were collected during the medits trawl surveys carried out in the sea of sardinia (2009-2014), during which environmental data were obtained with a multiparametric probe. a longitudinal (west-east) trend was found, with higher abundances at depths of 400-600 m, corresponding to salinity values of 38.1-38.5 psu and temperatures of 13.6-13.8°c. our results confirm the existence of a tight linkage between the distribution of the levantine intermediate water (liw) from the eastern mediterranean sea and the preferential habitat characteristics of the giant red shrimp. we suggest that a deeper knowledge of the relationships between abiotic (hydrological) factors in the water column and the distribution of mediterranean resources, such as the giant red shrimp, can provide valuable support for their better management, at the local scale (sardinia) and across the whole mediterranean sea. no nco mm er cia l u se on ly relationships between distribution of giant red shrimp and environmental factors 85 and seamounts) determine local hydrographic conditions that can facilitate species movement, thus contributing to the peculiar spatial distributions observed in this area (orrù and ulzega, 1988; sabatini et al., 2007). data come from the medits survey programme, an international bottom trawl survey, carried out in the mediterranean since 1994, every year, during the spring and the beginning of summer. this survey has been designed to sample all trawlable areas along coasts from 10 to 800 m depth. the application of a common standardized protocol allowed to produce biological data on demersal resources (spedicato et al., 2019). according to the medits protocol, the seas around sardinia were subdivided into seven zones: two located in the eastern coast (se-area 1, ne-area 2), one in the northern (n-area 3), three in the western coast (nw-area 4, cw-area 5, sw-area 6) and one in the southern coast (s-area 7) (fig. 1). data collection fishing data were collected from the sea of sardinia during the summer months from 2009 to 2014; trawl surveys were performed according to bertrand et al., 2002 using a stratified random sampling design, with the number of experimental hauls proportional to the surface of each depth stratum. a total of 103 experimental hauls were analysed (at depths of 400-700 m), corresponding to the sets in which potentially giant shrimp are found. trawl sensors (simrad) were connected to the net mouth to record the functioning and opening of the wings. data about horizontal opening net were used to evaluate the swept area (sparre and venema, 1998) and to compute standardized density index (di: number of individuals km–2). to investigate the effect of hydrological conditions on species abundance, data of both vertical and longitudinal profiles of temperature (°c), salinity (psu) and depth (m) were recorded using a multi-parameter probe ctd (sbe37 im microcat) mounted on the experimental net (in the otter of the trawler, goc73 net). for each longitudinal profile, we calculated the average temperature (bot_temp) and salinity (bot_sal) values at the bottom. for each vertical profile, we compiled the temperature (liwct) and salinity (liwcs) at the liw (levantine intermediate water) core, between 250 and 500 m, the average depth at the bottom (depth), the longitude and latitude coordinates (lat, lon), the spatio-temporal variables (year and sampling areas, areas 1-7) and fishing effort (n.a: number of bottom trawlers that operate in the sampling areas) (tab. 1). data on the trawling fleets active in the main ports of sardinia for the period 2009–2014 were obtained from the european fleet register (https://webgate.ec.europa.eu/fleet-europa/search_en). statistical analyses for the environmental variables, the zuur et al. (2010) protocol was followed, whereby collinearity was examined by computing pairwise scatter plots to compare continuous covariates; combinations with relevant spearman’s rho coefficients (ρ>0.7) were discarded prior to modelling. data exploration revealed non-linear patterns among the response variables, as such, giant red shrimp abundance and its relationship with environmental and spatio-temporal variables were described using generalized additive models (gams) (hastie and tibshirani, 1990; maunder and punt, 2004) and regression trees (walsh et al., 2001). gams (hastie and tibshirani, 1990; wood, 2006) are fig. 1. map of the study area. tab. 1. variables and acronyms used for the analysis. variable name of the variables bottom temperature bot_temp bottom salinity bot_sal liw core temperature liwct liw core salinity liwcs depth depth latitude lat longitude lon year of sampling year sampling areas areas fishing effort n.a no nco mm er cia l u se on ly c. podda et al.86 non-parametric regressions in which part of the linear predictor is specified as a sum of the smoothing functions (smooth function, s) of the predictor variables; the challenge is to find suitable parametric representations for the smoothing functions and appropriately control the degree of smoothness (wood and augustin, 2002). a stepwise backward selection procedure was implemented to identify the best fitting model, based on minimizing the akaike’s information criterion (aic) (akaike, 1973) values. model performances were evaluated by obtaining the total explained deviance. further approaches were based on the representation of generalized additive mixed models (gamms) as an extension of gams; gamms suggest a more complex structure than ordinary additive model and include smoothing terms as random effects (wood, 2006). the prediction of the spatial aggregations of species was obtained by means of gaussian process kriging model implemented in mgcv r package (wood, 2006). regression trees (morgan and sonquist, 1963), based on a recursive partitioning regression, were used to validate the obtained results. these models break data into left and right branches, whereby the splitting rules are defined by the predictor variable values. splitting continues until the ‘terminal’ nodes, wherein response values become similar within a node or data are too sparse for additional splitting. at the terminal node, the predicted response is the average or majority of the response values within that node for continuous or discrete variables. the sizes of the regression tree structure were examined because the vertical position of the nodes is an important function that reflects the degree of the relationship between variables (clark and pregibon, 1992). all statistical analyses were performed using r 3.3.1, with a significance level of p<0.05 (r core team, 2019). the gam approach, as proposed by wood (2006), was performed using the library mgcv and the regression tree with rpart. results density of giant red shrimp, sampled between 399 and 711 m depth, showed marked inter-annual fluctuations in all investigated areas. highest average densities were recorded in the southern (s-area 7), northern (n-area 3) and north-eastern areas of sardinia (ne-area 2), while the lowest density was recorded in north western area (nw-area 4) (tab. 2). the exploratory analysis of environmental data showed a significant correlation (ρ=0.93) between bottom salinity (bot_sal) and salinity at liw strata (liwcs). as such, bot_sal and liwcs were considered redundant and were then included separately in the predictive models. the best density model for a. foliacea was a gamm (aic=346.9) without random effect, that explains the 25% of the total deviance. the final model included geographical coordinates (lat, lon), depth (depth), bottom temperature (bot_temp), bottom salinity (bot_sal) and fishing effort (n.a) (tab. 3), according to the equation: (1) (di ~ s(lat, lon) + s(depth) + s(bot_temp) + s(bot_sal) + s(lat, lon, n.a)) (eq. 1) the cumulative effect of the covariates are illustrated in fig. 2, where the giant red shrimps’ aggregation areas are reported. highest abundances were recorded in southeastern and northern areas (area 3 and 7). important areas of aggregation were also observed throughout the eastern areas (areas 1, 2), where a. foliacea showed intermediate densities. depth, bottom temperature (bot_temp) and bottom salinity (bot_sal) showed a negative correlation. species density increased at depths between 400 and 600 m where bottom temperatures reach values between 13.6 °c and 13.8 °c and bottom salinity between 38.1 and 38.55 psu (fig. 2). the regression tree showed similar result with a significant relationship with environmental variables as selected in gamm models. the density was mainly influenced, at its first branch, by longitude values ≥4278000. later, a secondary branch was observed at depths ≥580.2 m (94 record). this node splits into two branches, which were influenced by latitude values ≥460000 (50 records) and latitude values <600000 (44 records), respectively, and average depth <565.6 m (20 records) and bottom salinity ≥38.55 psu, respectively (24 tab. 2. density index (average value ± se) of the investigated species in each one of the seven zones selected and each year. 2009 2010 2011 2012 2013 2014 area 1 639.4±27.6 2239.2±40.3 431.5±30.9 509±28.5 404.3±24.2 71.5±10 area 2 1053.75±41.8 178.2±18.2 758.3±34.7 10.5±4.6 199.7±15.4 455.8±22.8 area 3 1009±29.1 499±13.1 364.7±11.7 1965.7±45.4 1190±21.3 0 area 4 0 23±0 1364±0 1540±0 106±0 2736±0 area 5 78.5±4.1 231±0 12±4.9 0 870±46.9 104.5±14.4 area 6 256±0 148±0 315±0 115±0 24±1.2 103±0 area 7 2409.3±67.4 1646.8±59.6 202±0 1621.3±52.9 1013±38.2 800.6±44.2 no nco mm er cia l u se on ly relationships between distribution of giant red shrimp and environmental factors 87 records). our data confirm the significance of the results obtained from the gamm model: density was affected by the depth, longitude, latitude and bottom salinity (fig. 3). discussion and conclusions changes in environmental conditions can influence the life traits of marine organisms in different ways: by impacting spawning, growth and recruitment (pankhurst and munday, 2011; beggs et al., 2013), prey availability and prey-predator relationships (fanelli and cartes, 2010) or by altering their spatial distribution (perry et al., 2005). these impacts can be exacerbated by the effects of human activities, such as fishing pressure, which may impair the resistance and resilience of marine populations against environmental changes (anderson et al., 2008). due to the importance of a. foliacea in the deepwater communities of the mediterranean sea, many studies focused on the biology, ecology and fishery (palmas et al., 2017a and references therein). other authors have analysed the effects of different hydrological conditions (i.e., depth, temperature and salinity) on the species’ distribution (yahiaoui, 1994; cau et al., 2002; politou et al., 2004; sardà et al., 2004; company et al., 2008; eumofa, 2019). spatial distribution has also been related to the cascading of dense shelf waters along the slope (company et al., 2008), geomorphology (presence tab. 3. gamm summary results for the abundance of giant red shrimps. variables p-value % of variation explained s(lat, lon) 0.000565*** 25% s(detph) 3.62e-06*** s(bot_temp) 0.003380** s(bot_sal) 0.019274* s(lat, lon, n/a) 1.86e-05*** fig. 2. generalized additive model (gam)-derived effect of covariate modelling for the density index of giant red shrimps. shaded areas and dashed lines indicate 95% confidence bands. no nco mm er cia l u se on ly c. podda et al.88 of canyons and seamounts) (sabatini et al., 2007, 2011), bottom type (cartes et al., 2008), oceanographic features (guijarro et al., 2008) and fishing activity (d’onghia et al., 2005; carlucci et al., 2006). for instance, the abundance of the giant red shrimp would result high patchy as a consequence of several intermingling factors (rinelli et al., 2013, masnadi et al., 2018), which could act as drivers in shaping the spatial distribution of the species. the available literature reports a longitudinal gradient of the spatial distribution of a. foliacea, with abundances in the central and easternmost areas higher than those in the westernmost areas of the mediterranean sea (cau et al., 2002; politou et al., 2004; cardinale et al., 2017). in the mediterranean sea, giant red shrimp hotspots were observed: i) in the sardinian sea, where the species was more abundant in the southern grounds (cau et al., 2002); ii) in the central tyrrhenian sea, where the species reach greater numbers in the southern sector; iii) in southern sicily and in the western sector of the ionian sea where this species showed a consistent southern aggregation (ragonese et al., 1994; d’onghia et al., 2003). giant red shrimps occur at depths of 160-1330 m, and are prevalently caught at depths of 500–800 m (maiorano et al., 2010; bayhan et al., 2015; deval et al., 2016; eumofa, 2019, guijarro et al., 2019), although populations can be found also at shallower depths, typically between 100 and 160 m in the the ionian sea and in southern italy canyons (relini and relini-orsi, 1987; matarrese et al., 1995; d’onghia et al., 1996, sabatini et al., 2007). such a wide vertical distribution is due to the fact that the species is capable of wide daily movements along the water column (kapiris et al., 2010; fernández et al., 2013), ascending to shallower depths during the night (cau and deiana, 1982), particularly in winter. this phenomenon is more evident in canyons and seamounts, where bottom climbing on the continental slope edge can be observed (matarrese et al., 1995, sabatini et al., 2007, 2011; palmas et al., 2015, 2017b). a. foliacea prefers the warmer and more saline waters of the eastern mediterranean basin than the western basin. in particular, the species appreciates seawater temperatures close to 13°c and salinities of 38.5 psu (ghidalia and bourgois, 1961; yahiaoui, 1994; cartes et al., 2002; politou et al., 2004; sardà et al., 2004; company et al., 2008; noël, 2015), corresponding to the typical hydrological features of the levantine intermediate waters (liw) arriving from the eastern mediterranean basin. in detail, for the catalan sea and the balearic islands, red shrimps are abundant between 12.8 and 13.9°c. nevertheless, peak densities occur fig. 3. regression tree model on density index of a. foliacea. no nco mm er cia l u se on ly relationships between distribution of giant red shrimp and environmental factors 89 at ca. 12.8°c at depths between 900 and 1000 m (demestre and martín, 1993, sardà et al., 1998, tudela et al., 2003). this species can be found also from 80 to 600 m depth off algeria and tunisia at temperatures ranging from 12.8 to 14°c (yahiaoui, 1994). in the ionian sea, red shrimps have been reported at different depths, but the highest abundances were found at 600-800 m depths (deval et al., 2016), both in the western (at 13.3 and 13.7°c) and the eastern (at up 13.9°c) basin (politou et al., 2004). the hypothetical distribution range of this species could extend down to 2800 m depth (sardà et al., 2004). the spatio-temporal variability of the species abundance would seem also related to large-scale climatic indices, such as the north atlantic oscillation (nao) (d’onghia et al., 2012) even if the variability can differ among even nearby ports (hidalgo et al., 2015). all of the above delineated abiotic constraints are associated with an intense and prolonged fisheries exploitation, resulting in concurring effects which make difficult the interpretation of the whole picture of the species’ distribution (rinelli et al., 2013; sabatini et al., 2013). overall, our study confirms either the general assumption by which the abundance of a. foliacea follows a longitudinal eastern-western gradient or the influence of environmental variables in its spatial distribution and abundance in the seas surrounding sardinia (murenu et al., 1994; cau et al., 2002; rinelli et al., 2013). the highest abundances of giant red shrimp were observed in the southern (s-area 7) and northern (n-area 3) areas, confirming a longitudinal trend for the distribution of the species, with an increasing western-eastern pattern (guijarro et al., 2019). the particular variability in the hydrographic conditions of the sardinian seas determine the presence of different habitats that provide a complex system of environmental patches, which, in turn, are reflected in the distribution and abundance of the deep-water red shrimps along the sardinian slopes. our data confirm also that giant red shrimp, preferring depths of 400-600 m, salinity levels between 38.1 and 38.5 psu and temperatures between 13.6°c and 13.8°c, seem to concentrate in levantine intermediate waters (liw). in conclusion, due to response complexity, it is not always easy to establish unique relationships between a single environmental (abiotic or biotic) factor and a biological response. the multitude of pathways through which hydrological features affect marine populations often makes it difficult to establish univocal, significant and non-spurious connections between the climate and ecological responses (ottersen et al., 2010). in this work, the use of in situ environmental observations helped to clarify the role of some key environmental process on giant red shrimp abundance that can be extend across the whole mediterranean distributive scenario. corresponding author: cpodda@unica.it key words: aristaeomorpha foliacea; giant red shrimp; abundance; distribution models; environmental effects; mediterranean sea. received: 7 november 2020. accepted: 15 december 2020. this work is licensed under a creative commons attribution noncommercial 4.0 license (cc by-nc 4.0). ©copyright: the author(s), 2020 licensee pagepress, italy advances in oceanography and limnology, 2020; 11:9271 doi: 10.4081/aiol.2020.9471 references abelló p, carbonell a, torres p, 2002. biogeography of epibenthic crustaceans on the shelf and upper slope off the iberian peninsula mediterranean coasts: implications for the establishment of natural management areas. sci. mar. 66:183-198. akaike h, 1973. information theory and an extension of the maximum likelihood principle, p. 267-281. in: second international symposium on information theory. b.n. petrov, f. csaki (eds) budapest akademiai kiado. anderson cnk, hsieh c, sandin sa, hewitt r, hollowed a, beddington j, may rm, sugihara g., 2008. why fishing magnifies fluctuations in fish abundance. nature 452: 835-839. bayhan ky, cartes je, fanelli e, 2015. biological condition and trophic ecology of the deep-water shrimp aristaeomorpha foliacea in the levantine sea (sw turkey). mediterr. mar. sci. 16:46. beggs se, cardinale m, gowen rj, bartolino v, 2013. linking cod (gadus morhua) and climate: investigating variability in irish sea cod recruitment. fish. oceanogr. 23:54-64. bertrand ja, gil de sola l, papaconstantinou c, relini g, souplet a, 2002. the general specifications of the medits surveys. sci. mar. 66:9-17. bianchini ml, ragonese s, 1994. proceedings of the international workshop on ‘life cycles and fisheries of the deep water red shrimps aristaeomorpha foliacea and aristeus antennatus’. n.t.r. i.t.p.p. special publication 3: 87 pp. blanchard f, 2001. the effect of fishing on demersal fish community dynamics: an hypothesis. ices j. mar. sci. 58: 711-718. canals m, danovaro r, heussner s, lykousis v, puig p, trincardi f, calafat am, durrieu de madron x, palanques a, sanchezvidal a, 2009. cascades in mediterranean submarine grand canyon. oceanogr. 22:26-43. cardinale m, osio gc, scarcella g, 2017. mediterranean sea: a failure of the european fisheries management system. front. mar. sci. 4:72. carlucci r, d’onghia g, sion l, maiorano p, tursi a, 2006. selectivity parameters and size at first maturity in deep-water shrimps, aristaeomorpha foliacea (risso, 1827) and aristeus antennatus (risso, 1816), from the north-western ionian sea (mediterranean sea). in: thessalou-legaki, m. (ed.), issues no nco mm er cia l u se on ly c. podda et al.90 of decapod crustacean biology. hydrobiologia 557:145-154. carney rs, 2005. zonation of deep biota on continental margins. oceanogr. mar. biol. 43:211-278. cartes je, abelló p, lloris d, carbonell a, torres p, maynou f, gil de sola l, 2002. feeding guilds of western mediterranean demersal fish and crustaceans: an analysis based on a spring survey. sci. mar. 66:s209-220. cartes je, papiol v, guijarro b, 2008. the feeding and diet of the deep-sea shrimp aristeus antennatus off the balearic islands (western mediterranean): influence of environmental factors and relationship with the biological cycle. progress oceanogr. 79:37-54. cartes je, maynou f, fanelli e, 2011a. nile damming as plausible cause of extinction and drop in abundance of deep-sea shrimp in the western mediterranean over broad spatial scales. progress oceanogr. 91:286-294. cartes je, maynou f, abelló p, emelianov m, gil de sola l, solé m, 2011b. long-term changes in the abundance and deepening of the deep-sea shrimp aristaeomorpha foliacea in the balearic basin: relationships with hydrographic changes at the levantine intermediate water. j. mar. syst. 88:516-525. cau a, deiana am, 1982. [sulle variazioni di cattura del gambero rosso aristaeomorpha foliacea in relazione alla sua ecoetologia].[article in italian]. boll musei istit biol università genova 50:145-150. cau a, sabatini a, murenu m, follesa mc, cuccu d, 1994. [considerazioni sullo stato di sfruttamento delle risorse demersali (mari di sardegna) ].[article in italian]. biol. mar. mediterr. 1:67-76. cau a, carbonell a, follesa mc, mannini a, norrito g, orsi-relini l, politou cy, ragonese s, rinelli p, 2002. meditsbased information on the deep-water red shrimps aristaeomorpha foliacea and aristeus antennatus (crustacea: decapoda: aristeidae). sci. mar. 66:103-124. clark la, pregibon d, 1992. tree-based models p. 377-420. in: hastie jmcatj, editor: wadsworth. company jb, puig p, sardà f, palanques a, latasa m, scharek r, 2008. climate influence on deep sea populations. plos one 3:e1431. d’onghia g, mastrototaro f, matarrese a, politou cy, 2003. biodiversity of the upper slope demersal community in the eastern mediterranean: preliminary comparison between two areas with and without trawl fishing. j. northw. atl. fish. sci. 31:263-273. d’onghia g, capezzuto f, mytilineou c, maiorano p, kapiris k, carlucci r, sion l, tursi a, 2005. comparison of the population structure and dynamics of aristeus antennatus (risso, 1816) between exploited and unexploited areas in the mediterranean sea. fish. res. 76:22-38. d’onghia g, giove a, maiorano p, carlucci r, minerva m, capezzuto f, sion l, 2012. exploring relationshipsbetween demersal resources and environmental factors in the ionian sea (central mediterranean). j. mar. biol. 27946. demestre m, martìn p, 1993. optimum exploitation of a demersal resouce in the western mediterranean: the fishery of the deepwater shrimp aristeus antennatus (risso, 1816). sci. mar. 2:175-182. deval mc, kapiris k, 2016. a review of biological patterns of the blue-red shrimp aristeus antennatus in the mediterranean sea: a case study of the population of antalya bay, eastern mediterranean sea. sci. mar. 80:339-348. european market observatory for fisheries and aquaculture (eumofa), 2019. the eu fish market, 107 pp. available from: https://www.eumofa.eu/documents/20178/157549 /en_the+eu+fish+market_2019.pdf fanelli e, cartes je, 2010. temporal variations in the feeding habits and trophic levels of three deep-sea demersal fishes from the western mediterranean sea, based on stomach contents and stable isotope analyses. mar. ecol. prog. ser. 402:213-232. fernández mv, heras s, viñas j, maltagliati f, roldán mi, 2013. multilocus comparative phylogeography of two aristeid shrimps of high commercial interest (aristeus antennatus and aristaeomorpha foliacea) reveals different responses to past environmental changes. plos one 8(3): e59033. ghidalia w, bourgois f, 1961. [influence de la témperature et de l’éclairement sur la distribution des crevettes des moyennes et grandes profondeurs]. stud. rev. gen. fish. couns. medit. fao 16:1-53. guijarro b, massutí e, moranta j, díaz p, 2008. population dynamics of the red shrimp aristeus antennatus in the balearic islands (western mediterranean): short spatio-temporal differences and influence of environmental factors. j. mar. syst. 71:385-402. guijarro b, bitetto i, d’onghia g, follesa mc, kapiris k, mannini a, marković o, micallef r, ragonese s, skarvelis k, cau a, 2019. spatial and temporal patterns in the mediterranean populations of aristaeomorpha foliacea and aristeus antennatus (crustacea: decapoda: aristeidae) based on the medits surveys. sci. mar. 83:s57-70. guillen j, maynou f, floros c, sampsom d, conides a, kapiris k, 2012. a bioeconomic evaluation of the potential for establishing a commercial fishery on two newly developed stocks: the ionian red shrimp fishery. sci. mar. 76:597-605. hastie t, tibshirani r, 1990. generalized additive models. london: chapman and hall. hidalgo m, rueda l, molinero jc, guijarro b, massutì e, 2015. spatial and temporal variation of seasonal synchrony in the deep-sea shrimp aristeus antennatus in the western mediterranean. j. mar. syst. 148:131-141. kapiris k, thessalou-legaki m, petrakis g, conides a, 2010. ontogenetic shifts and temporal changes in the trophic patterns of the deep-sea red shrimp, aristaeomorpha foliacea (decapods: aristeidae), in the eastern ionian sea (eastern mediterranean). mar. ecol. 31:341-354. maiorano p, sion l, carlucci r, capezzuto f, giove a, costantino g, panza m, d’onghia g, tursi a, 2010. the demersal faunal assemblage of the north-western ionian sea (central mediterranean): current knowledge and perspectives. chem. ecol. 26:219-240. masnadi f, criscoli a, lanteri l, mannini a, osio gc, sartor p, sbrana m, ligas a, 2018. effects of environmental and anthropogenic drivers on the spatial distribution of deep-sea shrimps in the ligurian and tyrrhenian seas (nw mediterranean). hydrobiologia 816:165-178. massutì e, monserrat s, oliver p, moranta j, lópez-jurado jj, marcos m, hidalgo m, guijarro b, carbonell a, pereda p, 2008. the influence of oceanographic scenarios on the popno nco mm er cia l u se on ly relationships between distribution of giant red shrimp and environmental factors 91 ulation dynamics of demersal resources in the western mediterranean: hypothesis for hake and red shrimp off balearic islands. j. mar. syst. 71:421-438. matarrese a, d’onghia g, deflorio m, panza m, costantino g, 1995. [recenti acquisizioni sulla distribuzione batimetrica di aristaeomorpha foliacea e aristeus antennatus (crustacea, decapoda) nel mar ionio].[article in italian]. biol. mar. medit. 2:299-300. maunder mn, punt ae, 2004. standardizing catch and effort data: a review of recent approaches. fish. res. 70:141-159. morgan jn, sonquist ja, 1963. problems in the analysis of survey data, and a proposal. j. amer. statist. assoc. 58:415-434. murenu m, cuccu d, follesa c, sabatini a, cau a, 1994. the occurrence of aristaeomorpha foliacea in sardinian waters, p 49-50 in: life cycles and fisheries of the deep-water red shrimps aristaeomorpha foliacea and aristeus antennatus. proc. of the international workshop held in the instituto di tecnologia della pesca e del pescato. bianchini m. l., s. ragonese (editors), n.t.r. i.t.p.p. noël p, 2015. [la crevette rouge aristaeomorpha foliacea (risso, 1827)], p.1-6. [article in french]. in: muséum national d’histoire naturelle (ed.). inventaire national du patrimoine naturel. available from: https://inpn.mnhn.fr/fichesespece/especesmarines/aristaeomorpha_foliacea.pdf orrù p, ulzega a, 1988. [ricerche geomorfologiche sul canyon gonone (sardegna orientale].[article in italian]. bull. soc. r. sci. liège 57:415-427. ottersen g, kim s, huse g, polovina jj, stenseth nc, 2010. major pathways by which climate may force marine fish populations. j. marine syst. 79:343-360. palmas f, addis p, cabiddu s, cuccu d, follesa mc, mura m, olita a, pesci p, sabatini a, 2015. distribution of spawning and nursery grounds for deep–water red shrimps in the central western mediterranean sea. mediterr. mar. sci. 16:117-127. palmas f, maiorano p, sabatini a, 2017a. aristaeomorpha foliacea. in: sartor p., mannini a., carlucci r., massaro e., queirolo s., sabatini a., scarcella g., simoni r. (eds.), synthesis of the knowledge on biology, ecology and fishery of the halieutic resources of the italian seas. biol. mar. mediterr. 24:s1-11. palmas f, olita a, addis p, sorgente r, sabatini a, 2017b. modelling giant red shrimp larval dispersal in the sardinian seas: density and connectivity scenarios. fish. oceanogr. 26: 364-378. pankhurst nw, pl munday, 2011. effects of climate change on fish reproduction and early life history stages. mar. freshwater res. 62:1015-1026. papacostantinou c., kapiris k, 2003. the biology of the giant red shrimp (aristaeomorpha foliacea) at an unexploited fishing ground in the greek ionian sea. fish. res. 62:37-51. perry al, low pj, ellis jr, reynolds jd, 2005. climate change and distribution shifts in marine fishes. science 308: 1912-1915. politou cy, kapiris k, maiorano p, capezzuto f, dokos j, 2004. deep-sea mediterranean biology: the case of aristaeomorpha foliacea (risso, 1827) (crustacea: decapoda: aristeidae). sci. mar. 68:117-127. r core team, 2019. a language and environment for statistical computing (v. 3.6.0). r foundation for statistical computing, austria. ragonese s, bianchini ml, 1995. size at sexual maturity in red shrimp females, aristaeomorpha foliacea, from the sicilian channel (mediterranean sea). crustaceana 68:73-82. ragonese s, bianchini ml, di stefano l, 2002. trawl cod-end selectivity for deepwater red shrimp (aristaeomorpha foliacea, risso, 1827) in the strait of sicily (mediterranean sea). fish. res. 57:131-144. relini g, orsi relini l, 1987. the decline of red shrimp stocks in the gulf of genova. inv. pesq. 51:254-260. rinelli p, bianchini ml, casciaro l, giove a, mannini a, profeta a, politou cy, ragonese s, sabatini a, 2013. occurrence and abundance of the deep-water red shrimps aristeus antennatus (risso, 1816) and aristaeomorpha foliacea (risso, 1827) in the mediterranenan sea. cah. biol. mar. 54:335-347. sabatini a, follesa mc, locci i, pendugiu aa, pesci p, cau a, 2007. assemblages in a submarine canyon: influence of depth and time. hydrobiologia 580:265-271. sabatini a, follesa mc, locci i, matta g, palmas f, pendugiu aa, pesci p, cau a, 2011. demersal assemblages in two trawl fishing lanes located on the baronie seamount (central western mediterranean). j. mar. biol. assoc. u.k. 91:65-75. sabatini a, locci i, deiana am, follesa mc, gastoni a, pendungiu aa, pesci p, cau a, 2013. temporal trends in biodiversity of the middle-slope assemblages in sardinian seas (central-western mediterranean). j. mar. biol. assoc. u.k. 93:1739-1752. sala a, lucchetti a, perdichizzi a, herrmann b, rinelli p, 2015. is square-mesh better selective than larger mesh? a perspective on the management for mediterranean trawl fisheries. fish. res. 161:182-190. sardà f, maynou f, tallò l, 1998. seasonal and spatial mobility patterns of rose shrimp (aristeus antennatus, risso 1816) in the western mediterranean: results of a long-term study. mar. ecol. prog. ser. 159:133-141. sardà f, calafat a, mar flexas m, tselepides a, canals m, espino m, tursi a, 2004. an introduction to mediterranean deep-sea biology. sci. mar. 68:7-38. sbrana m, sartor p, belcari p, 2003. analysis of the factors affecting crustacean trawl fishery catch rates in the northern tyrrhenian sea (western mediterranean). fish. res. 65: 271-284. sparre p, venema sc, 1998. introduction to tropical fish stock assessment. part 1. fao fisheries technical paper 306/1. fao, rome: 337 pp. spedicato mt, massutí e, mérigot b, tserpes g, jadaud a, relini g, 2019. the medits trawl survey specifications in an ecosystem approach to fishery management. sci. mar. 83:9-20. stecf, 2015. mediterranean assessments part 1 (stecf-1518). eur 27638 en, jrc 98676. publications office of the european union, luxembourg. stecf, 2019. in: carvalho, n., keatinge, m., guillen garcia, j. (eds.), annual economic report on the eu fishing fleet (stecf 19-06). tudela s, sardà f, maynou f, demestre m, 2003. influence of submarine distribution of the deep-warer shrimp, aristeus antennatus (risso, 1816) in the nw mediterranean. crustaceana 76(2):217-225. walsh wa, kleiber p, 2001. generalized additive model and regression tree analyses of blue shark (prionace glauca) catch rates by the hawaii-based commercial longline fishery. fish. no nco mm er cia l u se on ly c. podda et al.92 res. 53:115-131. wood sn, augustin nh, 2002. gams with integrated model selection using penalized regression splines and applications to environmental modelling. ecol. model. 157:157-177. wood sn, 2006. generalized additive models: an introduction with r. crc. yahiaoui m, 1994. distribution and reproduction cycle of aristeus antennatus and aristeomorpha foliacea in algeria. in: bianchini m.l., ragonese s. (eds.), life cycles and fisheries of the deep-water red shrimps aristeomorpha foliacea and aristeus antennatus. proceedings of the international workshop, istituto di tecnologia della pesca e del pescato (itpp-cnr), mazara del vallo, 3:51-52. zuur af, ieno en, elphick cs, 2010. a protocol for data exploration to avoid common statistical problems. methods ecol. evol. 1:3-14. no nco mm er cia l u se on ly layout 1 introduction because of the growing anthropogenic impact, cyanobacterial blooms in freshwaters are increasing worldwide (huisman et al., 2018). eutrophication is the main driver of this phenomenon (conley et al., 2009), but climatic change also plays an important role in favoring the abnormal growth of cyanobacteria (mantzouki et al., 2018a; meriluoto et al., 2017a; moss et al., 2011; rigosi et al., 2014). many cyanobacterial species have the capability of producing toxic metabolites (cyanotoxins), like the hepatotoxic microcystins (mcs) and nodularins (nods), the cytotoxic cylindrospermopsins (cyns), the neurotoxic anatoxins (atxs) and saxitoxins (stxs) (carmichael, 1994; meriluoto et al., 2017b; metcalf and codd, 2012). the abundance of toxins during a cyanobacterial bloom is not directly and exclusively related to the biomass (mantzouki et al., 2018a; salmaso et al., 2014), but depends on several factors: first of all, the species which is involved in the bloom, as different species produce different toxins (cerasino et al., 2017; meriluoto et al., 2017b); and, secondly, the proportion of toxic strains within the cyanobacterial population (capelli et al., 2017; kurmayer et al., 2004). a third factor, represented by the environmental factors influencing the metabolic pathway of the toxin production, must be mentioned, although the mechanisms and the actual effects of abiotic and biotic factors are still a matter of debate (neilan et al., 2013). the presence of cyanotoxins in lakes can pose problems to water utilization and procedures for the management of the risks are regulated in many countries (chorus, 2012; ibelings et al., 2014). the italian perialpine district comprises five large and deep lakes (namely, lakes garda, iseo, como, lugano, and maggiore; dsl), which globally constitute over the 80% of the freshwater reserves in italy. dsl have experienced eutrophication in the 1960s and 1970s. the degradation of water quality was effectively contrasted by protection measures put in place since the 1980s. in the last two decades, nutrients levels (of phosphorous in particular) have declined consistently. together with eutrophication and re-oligotrophication, these lakes have been experiencing also common physical changes, represented by a significant warming and increased stability of the water column (anneville et al., 2005; gallina et al., 2013; salmaso et al., 2020, 2018b). both chemical and physical changes have deeply influenced the phytoplankton community. cyanobacteria have declined in recent years and their structure has changed. for instance, among toxic cyanobacteria, a decline of the mesotrophic planktothrix article co-occurrence of anatoxin-a and microcystins in lake garda and other deep perialpine lakes leonardo cerasino,* nico salmaso department of sustainable agro-ecosystems and bioresources, research and innovation centre, fondazione edmund mach, via e. mach 1, 38010 san michele all’adige, italy abstract cyanotoxins are a global concern in freshwaters and eutrophication and climate changes can have synergistic effects in exacerbating the problem. the deep perialpine lakes are a group of lakes of huge economic and naturalistic importance located at the border of the alps. at the southern border of the italian and swiss alps, the largest waterbodies include the lakes garda, iseo, como, lugano and maggiore (deep subalpine lakes, dsl). together with eutrophication (during the 1960s and 1970s) and re-oligotrophication (from the 1990s onward) these lakes have been experiencing warming and increase of the water column stability. these changes had a strong impact on the phytoplankton (including cyanobacteria) community. four dsl (lakes garda, iseo, como and lugano) have been studied with the aim of comparing their toxic potential. for one of them (lake garda) an 8 years survey was conducted, allowing a long-term trend analysis. toxin analysis was conducted on a monthly basis by targeted lc-ms/ms. a screening for anatoxins, cylindrospermopsins, saxitoxins, microcystins (mcs) and nodularins was carried out. among all the listed toxins, only one anatoxin and five mcs were detected in the lakes. in particular, the alkaloid anatoxin-a (atx) was found dominant in lakes garda, iseo and como, and absent in lake lugano; the mc-[d-asp3]rr was found as the most abundant mc in all four lakes. four other less abundant mcs were also found. the two major toxins are produced by two different cyanobacteria, tychonema bourrellyi (j.w.g. lund) anagnostidis & komárek and planktothrix rubescens (de candolle ex gomont) anagnostidis & komárek, which share however a number of ecological traits. peaks of these toxins occurred in warmer months (typically between may and september) in the thermocline layer (around 20 m, in the considered lakes). in summer 2016, the highest concentrations of atx and total mcs were registered in lake iseo (1100 and 430 ng l–1, respectively), while in the other lakes values were approximately twice lower. in the lakes where it was present, atx peak levels were much higher than mcs, thus highlighting the necessity of including atx in the procedures of risk assessment. the importance of atx is expected to further grow in the future with respect to mcs, as demonstrated by the long-term trend analysis carried out in lake garda that showed a clear decline for mcs from 2009 till 2016 and a relative constancy of atx. no nco mm er cia l u se on ly l. cerasino and n. salmaso12 rubescens (de candolle ex gomont) anagnostidis & komárek (mcs producer) was observed, which was partially replaced by tychonema bourrellyi (j.w.g.lund) anagnostidis & komárek (atxs producer). the two cited toxic species are part of the cyanobacterial populations of dsl, which also comprise microcystis aeruginosa (kützing) kützing, aphanizomenon flos-aquae ralfs ex bornet & flahault, and dolichospermum lemmermannii (richter) p.wacklin, l.hoffmann & j.komárek (cerasino et al., 2017; salmaso, 2019; salmaso et al., 2018a). analysis conducted on strains isolated form dsl (cerasino et al., 2017) have shown that m. aeruginosa and p. rubescens could produce mcs, t. bourrellyi anatoxin-a (atx), while a. flos-aquae and d. lemmermannii resulted to be not toxic. a successive study (capelli et al., 2017) showed also that the lack of toxicity of d. lemmermannii strains isolated in dsl was due to the lack of both mcs and atxs encoding genes. a field study was conducted in 2009 in order to evaluate the toxic potential of cyanobacterial populations in dsl (cerasino and salmaso, 2012); the study showed that i) mcs were present in all dsl, with a dominance of demethylated variants; and ii) atx was present in four out of five lakes, the exception being represented by lake lugano. moreover, in the study, the differences in toxin concentrations among lakes were interpreted in terms of different lake characteristics; more specifically, mcs levels were positively related to nutrients’ concentrations, while atx levels were more dependent on the water temperature, thus suggesting eutrophication and climate change as factors shaping the cyanotoxins diversity in different ways. climatic factors have been also recently reported as the main driver of cyanotoxins distribution at european level (mantzouki et al., 2018b). since 2009, cyanotoxins in lake garda have been regularly measured, thus allowing following the changes of the toxic potential of the lake. if the effects of the oligotrophication of lake garda and the other dsl is certainly protective against cyanobacteria proliferation, the effects of climate change are more difficult to assess, due to the many ways they can operate, for example increasing the water temperature and water column stability, increasing co2, and altering the hydrology (callieri et al., 2014; pareeth et al., 2017; visser et al., 2016). the availability of a time series for cyanotoxins data collected with uniform methodology is a valuable tool for understanding past changes and foreseeing future trajectories. this information is of great importance for the management of the lake, as the risks posed by mcs are different from those posed by atx, and also the measures for contrasting these risks are different (ibelings et al. 2014). moreover, considering the high chemical diversity of cyanotoxins, the analytical method should be adequate to describe as many toxins as possible. the objectives of this contribution include: i) critical evaluation of the results obtained from the analysis of cyanotoxins conducted with a monthly frequency in lake garda from 2009 till 2016; ii) the comparison of the longterm data collected in lake garda with those collected in other three dsl (lakes iseo, como and lugano) in 2016. data collected allow highlighting changes in the toxic profile in the considered lakes and contributing to define correctly the risks related to possible cyanobacterial blooms. methods sampling sites and sample collection the study was conducted in four lakes located in the italian perialpine region: garda, iseo, como and lugano. these lakes belong to the group of the dsl, which comprises also lake maggiore. the maximum depth of these lakes span between 251 m of lake iseo and 410 m of lake como, their altitude between 65 m of lake garda and 271 m of lake lugano, and their volume between 4.7×109 m3 of lake lugano and 49×109 m3 of lake garda. lakes garda, iseo and como host research stations belonging to lter (long term ecological research) italian network, and are part of the same macrosite (‘‘southern alpine lakes’’, it08-000-a, http://www.lteritalia.it/). all these lakes underwent an eutrophication process between the 1960s and 1980s (mosello et al., 1997); in the 1990s nutrient levels started to decrease and currently lakes garda and como can be classified as oligomesotrophic and lakes iseo and lugano as meso-eutrophic (mosello et al., 2010; salmaso et al., 2020). samples were taken with a monthly interval at the center of the lakes at discrete depths. lake garda was sampled from september 2009 until december 2016 (88 samplings) at four depths (0, 10, 20, 60 m); an additional sampling at 40 m was carried out only in 2016, in analogy with samplings carried out in the same year in lakes iseo, como and lugano, where samples were taken from five depths (0, 10, 20, 40, and 60 m). in 2016, lake iseo was sampled from february until december (12 samplings), lake como from april until december (10 samplings), and lake lugano only in june. at every depth, approx. 2.5 l of water were collected and stored in an amber glass bottle. in the lab, the water was filtered on 47 mm diameter gf-c filters (cytiva whatman), which were kept at -20°c until the analysis of toxins. toxin analysis lc-ms is the most powerful technique in cyanotoxins profiling thanks to the combination of high sensitivity and high selectivity (meriluoto et al., 2017b). the lc-ms methods used in this investigation have been setup through most of the standards currently available in the no nco mm er cia l u se on ly anatoxin-a and microcystins in perialpine lakes 13 market, which represent by far the most common and frequent. these methods have been applied previously on cultured strains of cyanobacteria taken from the lakes considered in this study and proved to be adequate to catch the toxin diversity in these environments (cerasino et al., 2017). filters were treated with 6 ml of acetonitrile/water mixture (60/40 v/v) containing 0.1% formic acid; the sample was first homogenized (omni th probe homogenizer, omni-inc.) for 5 min and then sonicated (omniruptor4000 probe sonicator, omni-inc.) for 4 min, using 160w power in pulsed mode (50%). the solution was separated from the pellet by centrifugation (eba 20, hettich) for 6 min at 9850 g. the pellet was then treated with a second aliquot of extraction mixture (6 ml) and sonicated. after centrifugation, the extraction liquid was added to the previous and vortexed. a 2 ml aliquot was filtered on 0.2 μm pore size rc syringe filters (phenomenex), and analyzed by lc-ms/ms. samples underwent two distinct lc-ms/ms analyses, one specific for peptides (mcs), and one specific for alkaloids (atxs, cyns, stxs). both analyses were performed using a waters acquity uplc system, directly coupled to a sciex 4000 qtrap hybrid mass spectrometer equipped with a turbo ion spray interface. mass detector was operated in scheduled mrm mode using positive electrospray ionization (esi+). details about protocols, instrumental settings and method performances can be found in cerasino et al. (2017, 2016). the analytical protocols were optimized for the detection and quantification of the peptidic toxins mcrr, [d-asp3]-rr, [d-asp3]-htyrr, yr, lr, [d-asp3]-lr, wr, la, ly, lw, lf, nod-r and of alkaloids atx, homoatx, cyn, saxitoxin (stx), decarbamoylstx, neostx, gtx 1, gtx4, gtx5, c1 and c2. these toxins were individually quantified in each sample. total mcs was calculated as the sum of each single mc variant (on a mass basis). cyanobacteria abundances from metagenomic data the relative abundances of t. bourrellyi and p. rubescens were obtained from the datasets published (ena study accession prjeb22244) and analyzed by salmaso et al. (2018a). the abundances of bacteria species (including cyanobacteria) were estimated as the number of reads after rarefaction of the otu table without replacement to 38000 sequences per sample. in this work, the abundances of tychonema and planktothrix were estimated as the % of reads on the total reads of cyanobacteria. statistical analysis data and graphical analyses were performed by using the r package 3.6.0 (r core team, 2019). the long-term trends of atx and mcs based on the data recorded in lake garda between 2009 and 2016 in the 0–20 m layer were evaluated using generalized additive models (gam), using the package mgcv in r (r core team, 2019), cfr. salmaso et al. (2016). the models were described by the effective degrees of freedom (edf); high edf values indicate higher non-linear smoothing spline (zuur et al., 2009). results toxins levels in lake garda in the 2009-2016 period in lake garda the presence of both toxin groups (mcs and atxs) was confirmed. total mcs and atx concentrations are graphically reported in fig. 1 a,b. seasonal toxin peaks, generally lasting a couple of months, were followed by longer periods (10 months, on average) in which toxins were present at concentrations typically two orders of magnitude lower. during the peaks, total mcs reached values higher than 100 ng l–1 in 2010, 2011, and 2015 (230, 200, and 190 ng l–1, respectively); in the remaining years, peak concentrations showed a decreasing trend from 90 (in 2009) to 20 ng l–1 (in 2016). on the other hand, for atx the peak concentrations showed values above 1000 ng l–1 from 2011 until 2014 (around 1300 in 2011 and 2012, 2700 in 2013, 2000 in 2014). a trend analysis, reported in fig. 1c, showed that, in the observed period, atx underwent an increasing tendency from 2009 until 2013, followed by a decreasing one in the following years. mcs (fig. 1d), instead, underwent a dramatic decrease: from higher values in the 2009-2011 period, concentrations dropped down to a minimum in the 2014-2016 period, with the remarkable exception of 2015. the annual toxin peaks occurred in different periods for atx and mcs. as shown in fig. 1 e,f, atx peaks typically were centered in june, while mcs peaks in august. the different temporal distribution of the two classes of toxins contributed to confirm that, in lake garda, they were produced by different organisms: atx by t. bourrellyi and mcs by p. rubescens. in fact, the comparison of toxin concentrations and cyanobacteria abundances derived from metagenomic data showed that that atx and mc levels were linearly correlated to t. bourrellyi and p. rubescens reads, respectively (fig. 2). fig. 3 a,b reports the vertical distributions of atx and mcs in lake garda. it clearly shows that the highest concentrations of both atx and mcs were present at 20 m depth. moreover, the figure shows that, during peaks, toxins are confined in a narrow layer around 20 m. for atx there was an early peak in april (900 ng l–1) and a second one, less intense, in june (500 ng l–1). for mcs, the peak was registered in september when a concentration of ca. 50 ng l–1. no nco mm er cia l u se on ly l. cerasino and n. salmaso14 fig. 1. data about cyanotoxin presence in lake garda from 2009 until 2016. monthly data about atx (a) and total mcs (b) concentrations in the lake (0-20 mean values). trend analysis for atx (c) and total mcs (d). seasonality of atx (e) and mcs (f). in c-f, the approximate significance of smooth terms (wood, 2006) is significant (p<0.05). no nco mm er cia l u se on ly anatoxin-a and microcystins in perialpine lakes 15 toxins levels in lakes iseo, como and lugano in 2016 toxin levels recorded for lakes iseo, como and lugano are graphically reported in fig. 3 c-f. similarly to lake garda, the analyses confirmed the presence of both atx and mcs in lakes iseo and como. in lake iseo (fig. 3 c,d), relevant concentrations of toxins were recorded from may till june, while in the rest of the year values were between 0 and 200 ng l–1 for atx and between 0 and 20 ng l–1 for total mcs. in may and june peak levels of atx (1300 and 1100 ng l–1) and total mcs (400 and 430 ng l–1) were reached at 20 m depth. differently from mcs, in june atx was spread toward the water surface, where a concentration of 560 ng l–1 was recorded. in lake como (fig. 3 e,f), we observed a similar pattern for toxin distribution, with highest values between june and august for atx and july and august for mcs. the highest concentrations of toxins were registered in june for atx (530 ng l–1) and in august for mcs (53 ng l–1) at the same depth of 20 m. in june, atx was uniformly present in the 0-20 m layer, and concentration at the lake surface was 300 ng l–1. in lake lugano, which was sampled only in june, mcs were found at levels of 2, 44, 17, 5 and 2 ng l–1 (as total mcs) at 0, 10, 20, 40 and 60 m depth, respectively. noteworthy, atx was not detected at all or in traces. toxin diversity while atx was the only alkaloid toxin found in the investigation, the peptidic toxins showed more diversity. in fact, five mc variants were found to be present, namely [d-asp3]-rr, [d-asp3]-lr, [d-asp3]-htyr, lr and rr. tab. 1 reports the abundances of these toxins in the different lakes. in all lakes, [d-asp3]-rr was always found as dominant over the others, with abundance always over 60% of the total mc content. the other variants were detected with lower frequency and lower abundancies: [d-asp3]fig. 2. relationship between toxins concentration and cyanobacterial abundances (estimated as fraction of reads in metagenomics analysis) in lake garda (2014-2015 only). both cyanobacteria abundances and toxins concentrations were double square transformed; relationships were highly significant (p<0.01). tab. 1. relative abundances of mc variants in the four lakes. for each variant the minimum and the maximum values recorded in the observation period are reported. lake mc variants (%) [d-asp3]-rr [d-asp3]-lr [d-asp3]-htyr lr rr garda 70 – 100 0 – 20 0 – 5 0 – 2 0 – 2 iseo 70 – 97 0 – 12 0 – 45 0 – 10 0 – 5 como 60 – 100 0 – 10 0 – 30 0 – 60 0 – 1 lugano 75 – 92 4 – 10 1 – 4 0 – 4 0 – 2 no nco mm er cia l u se on ly l. cerasino and n. salmaso16 fig. 3. vertical distribution of toxins measured in 2016 in lakes garda (a,b), iseo (c,d), and como (e,f). atx on the left hand side, total mcs on the right hand side. no nco mm er cia l u se on ly anatoxin-a and microcystins in perialpine lakes 17 lr represented the second abundant (between 0 and 20 %) in lake garda; this toxin reaches maximum 10-12% in the other lakes. [d-asp3]-htyr reached remarkable abundances in lakes iseo and como (45 and 30%, respectively), while it was always below 5 % in lakes garda and lugano. mc lr was always present in low abundances (below 4 %) in lakes garda and lugano, but reached 10 % in lake iseo and 60 % in lake como. finally, mc rr was found in very low abundances in all lakes (below 5%). in fig. 4, a comparative picture of the average mc diversity in the four lakes is reported. all lakes are dominated by the [d-asp3]-rr variant. in lake garda all other variants account for less than 8 %. [d-asp3]-htyr variant showed a higher contribution in lakes iseo and como (11 and 8.7 %, respectively), [d-asp3]-lr in lake lugano (8.7 %), and lr in lake iseo (6.9 %). discussion toxin dynamics in lake garda in the survey conducted in 2009 (cerasino and salmaso, 2012), mcs concentrations in lake garda were very low (maximum summer value was ca. 100 ng l–1), thus not representing a relevant issue. the survey for the first time revealed that another toxin, atx (an alkaloid with potent neurotoxic activity) was present in the lake with concentrations up to 5 times higher with respect to mcs: in 2009 a peak of 600 ng l–1 was found. this first observation triggered a series of investigations aimed at clarifying the origin of this toxin, finally leading to the identification of t. bourrellyi as the only atx producer in the lake (shams et al., 2015). this species seems to be fig. 4. averaged relative abundances (%) of different mc variants in the investigated lakes. no nco mm er cia l u se on ly l. cerasino and n. salmaso18 outcompeting the mcs producer p. rubescens in lake garda and in other dsl (salmaso et al., 2018b; salmaso et al., 2016). the data presented here (cfr. fig. 1), showed that in the eight years of observations, mcs and atx still continued to be both present in lake garda, with atx summer maxima always higher than mcs’. from 2011 till 2014, atx reached concentrations higher than 1 µg l–1 (the highest being 2.7 µgl–1 in 2013), which is considered a threshold limit for drinking water in some countries; mcs, instead, remained always below 0.25 µg l–1, abundantly lower than 1 µgl-1 set as threshold in drinking waters (ibelings et al., 2014). it is worth noting that toxin concentrations fell to very low values outside the warm season (may-october). mcs went often undetected in cold months and atx rarely exceeded 50 ng l–1. moreover, the toxin peaks were always limited to a narrow layer in the water columns, typically around 20 m, where producing organisms (t. bourrellyi and p. rubescens) find the best conditions to grow. it is worth noting that in all years atx and mcs summer peaks in lake garda are not synchronous. as reported in fig. 1, atx maxima generally occurred a couple of months earlier than mcs maxima, thus determining two distinct situations, an early one in which the cyanotoxins risk is due to atx and a late one linked to the prevalence of mcs. the evidence of this time separation can be helpful from a management point of view. finally, giving a global look at the data collected between 2009 and 2016 in lake garda, we can conclude that a decreasing trend is observable for both atx and mcs. this was more pronounced and clearer for mcs (fig. 1d) for which, excluding the exceptional 2015 year, a dramatic decline in summer concentrations was registered. for atx, on the other hand, the decline was evident only from 2014 onward, while there was a remarkable increase in previous years. these two toxins, although showing contrasting trends, globally seemed to be less and less present in the lake. toxins in the other dsl the results obtained from the 2016 sampling campaign in lakes iseo, como and lugano confirmed some evidences already obtained in 2009 (cerasino and salmaso, 2012). in fact, mcs were found in all the lakes, while atx was detected in lakes iseo and como, but not in lake lugano. the peak of toxin concentrations registered in lake iseo in 2016 was almost three times higher than those recorded in 2009 for both atx and mcs, when ca. 400 and 100 ng l–1, respectively, were found. in lake como, while mcs in 2016 were much lower than in 2009 (50 vs 200 ng l–1), atx levels remained similar (500 v. 440 ng l–1). finally, in lake lugano the mcs level showed a slight decrease from 100 to 44 ng l–1. differently from lake garda, peaks of atx and mcs in lakes iseo and como are less separated on a temporal scale (fig. 2). in lake como the two peaks are almost coincident (with maximum concentrations between may and june), while in lake iseo, they are only partially overlapped, with atx starting earlier (in june) and mcs a bit later (in july). the fact that for dsl we have data only from 2016 makes it impossible to affirm that this small temporal separation of atx and mcs is typical of these lakes. globally, this work suggests evidences of a decrease of mcs in lakes garda and como, a stable situation in lake lugano, and a definite increase of mcs in lake iseo. though based on a complete dataset only in lake garda, interestingly, these trends closely reflected the trophic status of the lakes (rogora et al., 2018; salmaso et al., 2020). lakes lugano and iseo, which registered constant concentrations or an increase of mcs, have higher nutrients content than lakes garda and como, which, on the contrary, experienced a decrease in mcs. these evidences strengthen the hypothesis of a correlation between mcs (and mcs producing organisms) with nutrient levels, which was formulated in cerasino and salmaso (2012). nevertheless, being based on a very limited set of data, the observations carried out in lakes iseo, como and lugano will have to be confirmed by extending the periods of investigations. on the other side, for atx, comparing 2009 and 2016, a stable situation was documented in lakes garda and como, and a net increase in lake iseo. overall, over eight years we observed a growing importance of atx at the expense of mcs in the dsl. interestingly, lake lugano was the only one among the dsl, in which no atx producing cyanobacteria have been described (capelli et al., 2018; capelli et al., 2017; shams et al., 2015). toxin diversity cyanobacteria can produce hundreds of different toxins. nearly 250 mcs and nods (spoof and arnaud, 2017), 5 cyns (wimmer et al., 2014), 10 atxs, 20 stxs (kellmann et al., 2013) are reported in the literature. this diversity is caused by the high flexibility of the biosynthetic pathways, which is typical of the secondary metabolism. it is common that each toxic species produces preferentially one group of toxins (atxs or mcs or stx) and it is also frequent that different variants of the same toxin group are produced by the same species (or strain). this happens more frequently in mcs producing cyanobacteria, in which the co-production of multiple variants is the rule. in alkaloids producing cyanobacteria, the variability is definitely lower. in experiments run on cultured cyanobacteria, we found that p. rubescens strains collected in lake garda produced five variants of mcs, but no alkaloids; t. bourrellyi, instead, produced only 2 alkaloids (atx and homoatx) and no no nco mm er cia l u se on ly anatoxin-a and microcystins in perialpine lakes 19 mcs; d. lemmermannii did produce neither mcs nor alkaloids (capelli et al., 2017; cerasino et al., 2017; salmaso et al., 2016). the mcs profile of cultured p. rubescens was coincident to the mcs profile found in lake garda in the 2009-2016 sampling campaign and in the other dsl lakes sampled in 2016, consisting in the presence of a dominant variant ([d-asp3]-rr, higher than 78%) and lower amounts of [d-asp3]-lr, [dasp3]-htyr, rr and lr. the dominance of demethylated (d-asp3) variants is widely considered as typical of p. rubescens (rohrlack et al., 2008; rounge et al., 2010). the other common mcs producer, also present in dsl, i.e. m. aeruginosa, has a mcs profiles in which standard non demethylated variants are dominant (namely, lr, yr and rr) (cerasino et al., 2017). while in lake garda the fraction of mc-lr was about 1 % of the total, in the other lakes this fraction was higher: 6.9% in lake iseo, 3.6% in lake como and 3.9% in lake lugano. these higher fractions of lr can be interpreted as a contribution of m. aeruginosa to the mcs pool, although there is the possibility that the differences among lakes is due to the presence of different chemotypes of p. rubescens. the constant prevalence of [d-asp3]-rr in all the investigated lakes should be taken into consideration for the assessment of the risk for humans and animals. currently, all the regulations are based on toxicological studies conducted on mc-lr. very little is known about the toxicity of the rr variant, which is also very frequent, and much less about the other hundreds of variants. although some studies have been conducted on rr variants (comprised the demethylated variants), results still do not allow deriving a conclusion about its higher or lower toxicity in comparison to the reference lr (buratti et al., 2017; diez-quijada et al., 2019). in relation to the presence of atx in dsl, it deserves particular attention because it is a potent neurotoxin that binds irreversibly to the acetylcholine receptors and blocks neurotransmission. atx produced by benthic cyanobacteria has been responsible of several animal fatalities (edwards et al., 1992; faassen et al., 2012; fastner et al., 2018). conclusions the investigation has allowed describing the evolution of the cyanotoxin diversity in lake garda from 2009 till 2016. the monthly frequency of sampling allowed detecting seasonal variations over a period of 8 years. in general, it was found that the alkaloid atx was the toxin with the highest concentrations, followed by [d-asp3]mcrr. four other mcs were commonly found but with much lower abundances. peaks of these toxins were generally observed in warmer months at depths between 20 and 30 m. the highest values were registered in 2013 for atx (2700 ng l–1) and 2010 for total mcs (230 ng l–1), both at 20 m. a trend analysis showed that both major toxins were subject to a decreasing trend, which was much more pronounced for mcs than atx. as a result, currently atx has assumed a noticeable relevance. lakes iseo and como have a situation very similar to lake garda: atx is found co-dominating together with [d-asp3]mc-rr. lake lugano differs from the others because no atx has been documented; the mcs diversity is, instead, very similar. compared to data collected in 2009, in 2016 atx and mcs concentrations in lake iseo were much higher, whereas, in the other lakes, toxins levels remained similar or lower. overall, data show that the issue of cyanotoxins in dsl needs to take into account the presence of mcs but must face the growing importance of atx. these toxins have different chemical and toxicological characteristics and therefore require different approaches in the risk management. acknowledgements investigations were carried out in the framework of the italian lter (long term ecological research) network, site southern alpine lakes, it08-000-a (http://www.lteritalia.it/), with the support of g. franzini and colleagues (arpa veneto), f. buzzi (arpa lombardia), and b. leoni (university of milan bicocca). special thanks are due to adriano boscaini and the technical staff in fem, for their support in the field and/or laboratory. corresponding author: leonardo.cerasino@fmach.it key words: cyanotoxins; anatoxin-a; microcystins; deep perialpine lakes; lake garda, lc-ms. received: 14 november 2019. accepted: 23 april 2020. this work is licensed under a creative commons attribution noncommercial 4.0 license (cc by-nc 4.0). ©copyright: the author(s), 2020 licensee pagepress, italy advances in oceanography and limnology, 2020; 11:8677 doi: 10.4081/aiol.2020.8677 references anneville o, gammeter s, straile d, 2005. phosphorus decrease and climate variability: mediators of synchrony in phytoplankton changes among european peri-alpine lakes. freshwater biol. 50:1731-1746. doi: 10.1111/j.13652427.2005.01429.x no nco mm er cia l u se on ly l. cerasino and n. salmaso20 buratti fm, manganelli m, vichi s, stefanelli m, scardala s, testai e, funari e, 2017. cyanotoxins: producing organisms, occurrence, toxicity, mechanism of action and human health toxicological risk evaluation. arch. toxicol. 91:1049-1130. doi: 10.1007/s00204-016-1913-6 callieri c, bertoni r, contesini m, bertoni f., 2014. lake level fluctuations boost toxic cyanobacterial “oligotrophic blooms”. plos one 9:8. doi: 10.1371/journal.pone.0109526 capelli c, ballot a, cerasino l, papini a, salmaso n, 2017. biogeography of bloom-forming microcystin producing and non-toxigenic populations of dolichospermum lemmermannii (cyanobacteria). harmful algae 67:1-12. doi: 10.1016/j.hal.2017.05.004 capelli c, cerasino l, boscaini a, salmaso n, 2018. molecular tools for the quantitative evaluation of potentially toxigenic tychonema bourrellyi (cyanobacteria, oscillatoriales) in large lakes. hydrobiologia 824:109-119. doi: 10.1007/s10750018-3513-3 carmichael ww, 1994. the toxins of cyanobacteria. sci. am. 270:78-86. doi: 10.1038/scientificamerican0194-78 cerasino l, capelli c, salmaso n, 2017. a comparative study of the metabolic profiles of common nuisance cyanobacteria in southern perialpine lakes. adv. ocean. limnol. 8:22-32. doi: 10.4081/aiol.2017.6381 cerasino l, salmaso n, 2012. diversity and distribution of cyanobacterial toxins in the italian subalpine lacustrine district. oceanol. hydrobiol. st. 41:54-63. doi: 10.2478/s13545-012-0028-9 cerasino l, shams s, boscaini a, salmaso n, 2016. multiannual trend of microcystin production in the toxic cyanobacterium planktothrix rubescens in lake garda (italy). chem. ecol. 32:492-506. doi: 10.1080/02757540.2016.1157175 chorus i, 2012. current approaches to cyanotoxin risk assessment, risk management and regulations in different countries. federal environment agency: dessau-rosslau, germany. conley dj, paerl hw, howarth rw, boesch df, seitzinger sp, havens ke, lancelot c, likens ge. 2009. ecology controlling eutrophication: nitrogen and phosphorus. science 323:1014-1015. doi: 10.1126/science.1167755 diez-quijada l, puerto m, gutierrez-praena d, llana-ruizcabello m, jos a, camean am, 2019. microcystin-rr: occurrence, content in water and food and toxicological studies. a review. environ. res. 168:467-489. doi: 10.1016/j.envres.2018.07.019 edwards c, beattie ka, scrimgeour cm, codd ga, 1992. identification of anatoxin-a in benthic cyanobacteria (bluegreen-algae) and in associated dog poisoning at loch insh, scotland. toxicon 30:1165-1175. doi: 10.1016/00410101(92)90432-5 faassen ej, harkema l, begeman l, lurling m, 2012. first report of (homo)anatoxin-a and dog neurotoxicosis after ingestion of benthic cyanobacteria in the netherlands. toxicon 60:378-384. doi: 10.1016/j.toxicon.2012.04.335 fastner j, beulker c, geiser b, hoffmann a, kroger r, teske k, hoppe j, mundhenk l, neurath h, sagebiel d, chorus i, 2018. fatal neurotoxicosis in dogs associated with tychoplanktic, anatoxin-a producing tychonema sp in mesotrophic lake tegel, berlin. toxins 10:11. doi: 10.3390/toxins10020060 gallina n, salmaso n, morabito g, beniston m, 2013. phytoplankton configuration in six deep lakes in the perialpine region: are the key drivers related to eutrophication and climate? aquat. ecol. 47:177-193. doi: 10.1007/s10452-013-9433-4 huisman j, codd ga, paerl hw, ibelings bw, verspagen jmh, visser pm, 2018. cyanobacterial blooms. nat. rev. microbiol. 16:471-483. doi: 10.1038/s41579-018-0040-1 ibelings bw, backer lc, kardinaal wea, chorus i, 2014. current approaches to cyanotoxin risk assessment and risk management around the globe. harmful algae 40:63-74. doi: 10.1016/j.hal.2014.10.002 kellmann r, ploux o, neilan ba, 2013. neurotoxic alkaloids from cyanobacteria. in: natural products. ramawatt kg, merillon jm, eds. springer-verlag, berlin heidelberg. doi: 10.1007/978-3-642-22144-6_47 kurmayer r, christiansen g, fastner j, börner t, 2004. abundance of active and inactive microcystin genotypes in populations of the toxic cyanobacterium planktothrix spp. environ. microbiol. 6:831-841. doi: 10.1111/j.14622920.2004.00626.x mantzouki e, campbell j, van loon e, visser p, konstantinou i, antoniou m, et al, 2018a. data descriptor: a european multi lake survey dataset of environmental variables, phytoplankton pigments and cyanotoxins. sci. data 5:180226. doi: 10.1038/sdata.2018.226 mantzouki e, lurling m, fastner j, domis ld, wilk-wozniak e, koreiviene j, et al, 2018b. temperature effects explain continental scale distribution of cyanobacterial toxins. toxins 10:156. doi: 10.3390/toxins10040156 meriluoto j, blaha l, bojadzija g, bormans m, brient l, codd ga, et al, 2017a. toxic cyanobacteria and cyanotoxins in european waters – recent progress achieved through the cyanocost action and challenges for further research. adv. oceanogr. limnol. 8:161-178. doi: 10.4081/aiol. 2017.6429 meriluoto j, spoof l, codd g, 2017b. handbook of cyanobacterial monitoring and cyanotoxin analysis. meriluoto j, spoof l, codd ga, eds.; wiley: chichester, uk metcalf js, codd ga, 2012. cyanotoxins. in: ecology of cyanobacteria ii. springer, dordrecht. mosello r, ambrosetti w, arisci s, bettinetti r, buzzi f, calderoni a, et al, 2010. evoluzione recente della qualità delle acque dei laghi profondi subalpini (maggiore, lugano, como, iseo e garda) in risposta alle pressioni antropiche e alle variazioni climatiche. biologia ambientale 24:167-177. mosello r, calderoni a, de bernardi r, 1997. [le indagini sulla evoluzione dei laghi profondi sudalpini svolte dal c.n.r. istituto italiano di idrobiologia].[article in italian]. doc. ist. ital. idrobiolog. 61:19-32. moss b, kosten s, meerhoff m, battarbee rw, jeppesen e, mazzeo n, et al, 2011. allied attack: climate change and eutrophication. inland waters 1:101-105. doi: 10.5268/iw1.2.359 neilan ba, pearson la, muenchhoff j, moffitt mc, dittmann e, 2013. environmental conditions that influence toxin biosynthesis in cyanobacteria. environ. microbiol. 15:12391253. doi: 10.1111/j.1462-2920.2012.02729.x pareeth s, bresciani m, buzzi f, leoni b, lepori f, ludovisi a, et al, 2017. warming trends of perialpine lakes from no nco mm er cia l u se on ly anatoxin-a and microcystins in perialpine lakes 21 homogenised time series of historical satellite and in-situ data. sci. total env. 578:417-426. doi: 10.1016/j.scitotenv. 2016.10.199 r core team, 2019. r: a language and environment for statistical computing (v. 3.6.0). r foundation for statistical computing, wien, austria. rigosi a, carey cc, ibelings bw, brookes jd, 2014. the interaction between climate warming and eutrophication to promote cyanobacteria is dependent on trophic state and varies among taxa. limnol. oceanogr. 59:99-114. doi: 10.4319/lo.2014.59.1.0099 rogora m, buzzi f, dresti c, leoni b, lepori f, mosello r, et al, 2018. climatic effects on vertical mixing and deep-water oxygen content in the subalpine lakes in italy. hydrobiologia 824:33-50. doi: 10.1007/s10750-018-3623-y rohrlack t, edvardsen b, skulberg r, halstvedt cb, utkilen hc, ptacnik r, skulberg om, 2008. oligopeptide chemotypes of the toxic freshwater cyanobacterium planktothrix can form subpopulations with dissimilar ecological traits. limnol. oceanogr. 53:1279-1293. doi: 10.4319/lo.2008.53.4.1279 rounge tb, rohrlack t, decenciere b, edvardsen b, kristensen t, jakobsen ks, 2010. subpopulations differentiation associated with nonribosomal peptide synthetase gene cluster dynamics in the cyanobacterium planktothrix spp. j. phycol. 46:645-652. doi: 10.1111/j.1529-8817.2010.00856.x salmaso n, buzzi f, capelli c, cerasino l, leoni b, lepori f, rogora m, 2020. responses to local and global stressors in the large southern perialpine lakes: present status and challenges for research and management. j. great lakes res. doi: 10.1016/j.jglr.2020.01.017 salmaso n, 2019. effects of habitat partitioning on the distribution of bacterioplankton in deep lakes. front. microbiol. 10. doi: 10.3389/fmicb.2019.02257 salmaso n, albanese d, capelli c, boscaini a, pindo m, donati c, 2018a. diversity and cyclical seasonal transitions in the bacterial community in a large and deep perialpine lake. microb. ecol. 76:125-143. doi: 10.1007/s00248-017-1120-x salmaso n, boscaini a, capelli c, cerasino l, 2018b. ongoing ecological shifts in a large lake are driven by climate change and eutrophication: evidences from a three-decade study in lake garda. hydrobiologia 824:177-195. doi: 10.1007/s10750-017-3402-1 salmaso n, cerasino l, boscaini a, capelli c, 2016. planktic tychonema (cyanobacteria) in the large lakes south of the alps: phylogenetic assessment and toxigenic potential. fems microbial ecol. 92. doi: 10.1093/femsec/fiw155 salmaso n, copetti d, cerasino l, shams s, capelli c, boscaini a, et al, 2014. variability of microcystin cell quota in metapopulations of planktothrix rubescens: causes and implications for water management. toxicon 90:82-96. doi: 10.1016/j.toxicon.2014.07.022 shams s, capelli c, cerasino l, ballot a, dietrich dr, sivonen k, salmaso n, 2015. anatoxin-a producing tychonema (cyanobacteria) in european waterbodies. water research 69:68-79. doi: 10.1016/j.watres.2014.11.006 spoof l, arnaud c, 2017. appendix 3: tables of microcystins and nodularins. in: meriluoto j, spoof l, codd ga (eds.), handbook of cyanobacterial monitoring and cyanotoxin analysis. wiley, chichester. visser pm, verspagen jmh, sandrini g, stal lj, matthijs hcp, davis tw, et al, 2016. how rising co2 and global warming may stimulate harmful cyanobacterial blooms. in harmful algae 54:145-159. doi: 10.1016/j.hal.2015.12.006 wimmer km, strangman wk, wright jlc, 2014. 7-deoxydesulfo-cylindrospermopsin and 7-deoxy-desulfo12-acetylcylindrospermopsin: two new cylindrospermopsin analogs isolated from a thai strain of cylindrospermopsis raciborskii. harmful algae 37:203-206. doi: 10.1016/j.hal.2014.06.006 wood, s, 2006. generalized additive models: an introduction with r. chapman and hall/crc. zuur af, ieno en, walker nj, saveliev aa, smith gm, 2009. mixed effects models and extensions in ecology with r. springer, new york. no nco mm er cia l u se on ly layout 1 introduction in aquatic ecosystems, microorganisms play a variety of important roles at the foundation of the microbial food webs as prey, predators or parasites, but also as actors in decomposing organic matter and (re)cycling nutrients. among such microorganisms, archaeal contribution to the microbial loop and the interactions involved between archaea and other prokaryotes have been partially examined so that many pieces of information still remain to be explored and discovered (braga et al., 2016; parada and fuhrman, 2017; seyler et al., 2019). for a long time, archaea were believed to only inhabit extreme environments. high temperature, salinity and extreme anaerobiosis conditions were thought to be obligate prerequisite for archaeal growth. therefore, possible links between archaea and other prokaryotes, typically bacteria, were not examined because not expected in such habitats where bacterial growth is improbable or even not possible (seyler et al., 2019). for the last two decades, however, it has been shown that archaea can be found everywhere (fuhrman and davis, 1997; robertson et al., 2005) including pelagic and surface freshwater ecosystems (hugoni et al., 2013; casamayor, 2017). along the last two decades, archaea were even reported to be as abundant as bacteria in marine sediments (wurzbacher et al., 2017) or dominant in deep ocean prokaryotic communities (varela et al., 2008). archaea, like bacteria, are important players in biogeochemical cycles, i.e. the carbon and nitrogen cycles. moreover, they are the only life forms to operate some fundamental processes such as methanogenesis (cavicchioli, 2011) which is a very important function of anaerobic lake waters and sediments (koizumi et al., 2004; bomberg et al., 2008). archaea are also known to oxidize ammonia (könneke et al., 2005), degrade ammonium urea and, like bacteria, to operate the nitrification process (alonso-sáez et al., 2012). operating comparable or same functions may logically generate a competition between archaea and bacteria. therefore, a variety of factors including nutrient availability is likely to structure differently these communities (berdjeb et al., 2011, 2013). this does not exclude that they can coexist in a non-limiting environment for growth with different responses to resource availability and/or predation pressure by viruses or planktonic grazers (seyler et al., 2019). it is noteworthy, however, that archaeal growth seems to be favored over bacteria in oligotrophic environments (vuillemin et al., 2019). still today, the relationships between archaea and bacteria remain largely unknown. in the microbial realm, we can find relationships such as mutualism, amensalism, competition or elimination for example through antibiotics secretion. therefore, the outcomes of such relations can be described as positive, neutral or negative (pacheco and segrè, 2019). for instance, we can cite the parasitic life style for pacearchaeota and woesearchaeota with bacteria (ortiz-alvarez and casamayor, 2016). another contact dependence has also been reported for two archaea, where nanoarchaeum equitans directly depends on igniococcus hospitalis for its survival (hu et al., 2018). this could be explained by the fact that some archaea have a compact genome (koonin and wolf, 2008) and limited metabolic capabilities (ortizalvarez and casamayor, 2016) so that they could need a partnership to compensate. another example is provided by some nitrogen-fixing archaea, designated as a single entity which is known to engage an interaction with the bacterial genus desulfosarcina (dekas et al., 2009). the aim of this study was to assess, for the first time using high throughput sequencing, the diversity of both bacterial and archaeal assemblages in lake geneva and to investigate the potential relationships between these two communities using co-occurrence network analysis and functional profile prediction. we tested the hypothesis that archaea and bacteria may share similar niches and display significant links for a variety of ecological or biogeochemical functions. article exploring archaeal and bacterial diversity and co-occurrence in lake geneva jade a. ezzedine,1 yves desdevises,2 stéphan jacquet1* 1inrae, umr carrtel, université savoie mont-blanc, thonon-les-bains; 2cnrs, biologie intégrative des organismes marins, observatoire océanologique, 2sorbonne université, f-66650 banyuls-sur-mer, france abstract the diversity and relationships between archaea and bacteria remain poorly examined in lakes. using universal primers targeting 16s rrna gene via hiseq sequencing, we explored archaeal and bacterial diversity, structure and relationships in the largest natural deep lake in western europe, i.e. lake geneva. despite being less diverse than bacteria, archaeal dominant otus assigned to the phylum thaumarchaeota and nanoarchaeota displayed significant links with a variety of nitrifying bacteria and other bacteria as suggested by co-occurrence networks and function profile predictions. we propose that archaeal otus are most likely involved in nitrogen and methanogen cycles, and formed a consortium with other bacteria that also operate these cycles in the deep layers of the lake. these probable syntrophic or mutualistic associations suggest that dominant archaeal otus share with some bacteria a similar niche for mutual benefits. j.a. ezzedine et al.46 methods study site lake geneva is a deep and large warm monomictic lake (surface area: 580 km2; volume: 89 km3; maximum depth: 309.7 m; mean depth: 152.7 m), located in the western part of the alps at an altitude of 372 m. lake geneva has been monitored since 1974 as a part of a long-term water quality and biological monitoring program. sampling has been continuously undertaken in the middle of the lake at the deepest point, referred to as shl2, once or twice a month. this scientific survey revealed that the lake has switched from an oligotrophic to eutrophic state with annual phosphorus concentrations reaching 90 μg l–1 in 1979 (anneville et al., 2002). thanks to effective management measures, lake geneva turned back to a mesotrophic state in the early 2000s with total phosphorus concentrations about 20 μg l–1 in 2010 (jacquet et al., 2014). sampling strategy during the translem project, we collected water samples at three sites, including the reference station shl2. these sites referred to as pt2, pt4 (shl2) and pt6 (supplementary fig. s1), separated from each other by approximately 14 km, were sampled at four different dates and seasons: february 20 (tl1), june 4 (tl2), august 7 (tl3) and november 20 (tl4) 2014, and at three different depths, i.e. 2, 15 and 200 m. it is noteworthy that pt2 could not be sampled on june 4 due to bad weather conditions. a filtration system was set on the boat and a volume of 300 ml of each water sample was filtered through 5 (to select all prokaryotes and avoid clogging of the 2µm filter), 2 and 0.2 µm polycarbonate 47 mm filters (merck millipore, burlington, ma, usa). samples were then kept at -20°c until dna extraction. descriptors such as temperature, ph, conductivity, chlorophyll a and dissolved oxygen concentrations of the water column were measured using a multiparametric probe (sea & sun technology gmbh, trappenkamp, germany). transparency was measured using a normalized 25 cm diameter secchi disk. total organic carbon (toc), total nitrogen (tn), dissolved ammonium (nh4n), dissolved nitrates (no3-n), total phosphorus (tp) and orthophosphates (po4-p) were measured only at pt4, at the different depths and dates, according to the standard french protocols afnor. dna extraction the filters were subjected to dna extraction using a homemade protocol with genelutetm-lpa (sigmaaldrich, st. louis, mo, usa) solution. the protocol started with a lysis step in eppendorf tubes by adding 300 µl of te buffer (tris 1m – ph 8, edta 0.5m – ph 8) and 200 µl of a lysis solution (tris 1m – ph 8, edta 0.5 m – ph 8 and sucrose 0.7 m). next, a thermic shock was carried out by first placing the tubes at -80°c for 15 min and then thawed into a block heater at 55°c for 2 min. after, 50 µl of a 10% sodium dodecyl sulfate (sds) and 10 µl of proteinase k (20 mg ml–1) were added to the solution. the solution was incubated at 37°c for 1 h with gentle stirring and placed again in the block heater at 55°c for 20 min. after a quick centrifugation step (13,000 rpm at 4°c for 3 min), the supernatant was collected. then, 50 µl of sodium acetate (3 m – ph 5.2) and 1 µl of genelutetm-lpa (sigma-aldrich, 25 µg µl–1) were added. one volume of isopropanol was then added and the tubes were centrifuged for 10 min at 12,000 g and 4°c. the supernatant was discarded and 2 washing rounds using ethanol (80%) were carried to purify the dna. the remaining ethanol was evaporated using a speed-vac for 20 min. finally, 30 µl of te were added and tubes were left for 1 h at 37°c. dna concentration was measured using a nanodrop 1000 spectrophotometer. the amount of dna extracted from 5 µm size filters ranged between 4.25 and 138.09 ng µl–1 with an average of 45.37 ng µl– 1. for the 2 µm filters, the minimum, maximum and mean concentrations were 6.71, 166.77, and 59.75 ng µl–1, respectively. for the 0.2 µm filters such minimal, maximal and mean concentrations were 2.93, 215.18 and 66.89 ng µl–1 respectively. afterwards, all tubes were stored at 20°c until analysis. pcr and sequencing total dna extracts were set at 25 ng µl–1. total dna extracted from the 5 and 2 µm filters were pooled since we had low amplification when considering only 2 µm filters. 0.2 µm filter was considered as it is. the pcr amplification of 16s rrna gene fragments was performed using universal set of primers. the primers, namely forward primer 515f (gtgycagcmgccgcggta) (wang and qian, 2009) and reverse primer 909r (ccccgycaattcmtttragt) (wang et al., 2018) had tags attached to them. it is noteworthy here that we checked, using tesprime (not shown), that this set of primers had a good coverage toward prokaryotes. the final number of samples was 66 and a pcr replicate was made for each sample giving a total of 132 samples. each sample was identified with a different tag. pcr mixture volume was 30 µl and consisted of (final concentration): 1x buffer, 0.5 mm dntp, 1.5 mm mgcl2, 0.5 mg ml–1 bovine serum albumin (bsa) and 0.75 u biotaq dna polymerase (bioline). in a second step, a unique combination of tagged primers (forward and reverse) was added to each sample. finally, 1 µl of template dna (25 ng µl– 1) was added. a negative control was included and the pcr program was as follows: 95°c – 2 min, 30 x (94°c lake microbial statistical interactions 47 – 30 sec, 58°c – 30 sec, 72°c – 30 sec), with a final extension step at 72°c for 5 min. agarose gel analysis was performed to check the pcr products. when samples showed a non-specific band (approximately 550 pb) close to the target band (approximately 450 pb), the target bands were then captured using pippin prep system (sage science) following the manufacturer’s instructions. these captured dna were checked with tapestation (agilent 2200) system for size and quality assessment following the manufacturer’s instructions. all amplified dna were quantified using the quant-it picogreen ds dna reagent kit (invitrogen) and fluorescence was read using the plate reader fluoroskan ascent fl. all dna samples were then pooled as one equimolar tube. dna was then purified using the clean pcr kit (cleanna) according to the manufacturer’s instructions to remove dntp and dimers. again, the pool was quantified using picogreen. then, the pool tube containing the 132 different tagged dna was sent to the gatc-eurofins platform for dna sequencing using illumina hiseq paired end technology to get 2 x 250 bp amplicons. bioinformatics pipeline two paired fastq files, referred to as r1 (forward sequences) and r2 (reverse sequences), were received from the sequencing platform. files were processed using the pipeline developed by frederic mahé which combines vsearch (rognes et al., 2016), cutadapt (martin, 2011) and swarm (mahé et al., 2015) and reported at https://github.com/frederic-mahe/swarm/wiki/fred’smetabarcoding-pipeline. all default parameters of the pipeline were left unchanged except when mentioned. briefly, the pipeline started with merging reads of r1 and r2 files using vsearch. next, cutadapt was used to demultiplex the sequences according to a list of tags. here, each sequence was assigned to its sample in an individual file by its tag. primers were also removed and sequences containing ambiguous bases were discarded. in each file, sequences were dereplicated using vsearch. all files were then assembled as one file and dereplicated in the process. the swarm algorithm was then used to cluster the sequences. the “d” parameter (i.e. the number of different nucleotides between sequences) was changed from 1 to 13 in order to be close to the identity threshold of 97% which describes same species. for the calculation, we took into account the median length of sequences, which is 377 bp. next, de novo chimera detection was applied to representative sequences of each cluster using vsearch. using stampa (sequence taxonomic assignment by massive pairwise alignments), the representative sequences were taxonomically assigned to a reference database downloaded from arb-silva (release number 132; (quast et al., 2013)) and prepared as required. the final otu table was built using the python script provided in the pipeline. then, multiple filtering was carried out on the otus table and these steps were named “default”, “stringent”, and “shared”. in the “default” step, the otu table was filtered by removing chimera, low quality sequences (<0.0002) and otus with less than 3 reads in a sample, unless they were present in 2 or more samples. after examining the data, we applied the “stringent” filter. it consisted of keeping otus with 10 or more reads present in 2 or more samples, removing otus with an identity score lower than 90% to arb-silva and discarding unassigned otus. finally, we applied the last filter, “shared”, where we kept otus that are common between the pcr replicates. on the other hand, we intended to study microorganisms attached to particles, i.e. those captured on the 2 µm filter vs the free-living part collected between 2 and 0.2 µm. however, due to technical issues, we chose to associate reads of all filters (5+2 µm and 0.2 µm), providing prokaryotes of all size. at the end, after applying all the aforementioned filters and association, only 33 out of 132 samples were analyzed. statistical analysis statistical analyses and plots were performed using r, version 3.5.0 (r core team, 2019) with the ggplot2 package (wickham et al., 2018). the otu table was transformed to relative abundance using the “decostand” function from the vegan package (oksanen et al., 2019). thus, all statistical analysis was performed using relative abundance. alpha diversity indices (i.e. shannon, pielou, chao1) were calculated with the otutable package (linz, 2018). simpson and inverse simpson were computed using the vegan package (oksanen et al., 2019). the indexes comparison for each condition (site, month, and depth) was performed using anova or kruskal-wallis test according to the data distribution after initial testing with shapiro and bartlett test. when the p-value of anova or kruskal-wallis test was inferior to 0.05 (alpha), a tukey hsd or a dunn test with bonferroni correction (alexis dinno, 2017) were performed. an nmds to illustrate beta diversity was computed using the “metamds” function from vegan and also the goeveg package (goral and schellenberg, 2017). also, adonis and anosim (oksanen et al., 2019) tests were performed to analyse whether the groups observed were significantly different. moreover, a simper test (oksanen et al., 2019) was performed to assess the contribution of each otu to the observed dissimilarity between samples. regarding environmental variables, they were only available for pt4. otus of pt4 were extracted and if an otu had no read in that site, the otu was removed from the final otus table. the raw data was transformed as log(1 + x) in order to stabilize variances, and converted to relative abundance in order to perform a cca following the online tutorial of umer zeeshan ijaz (torondel et al., 2016) found at j.a. ezzedine et al.48 http://userweb.eng.gla.ac.uk/umer.ijaz/bioinformatics/eco logical.html. redundant environmental variables were removed after calculating a variance inflation factor (vif). any environmental variable with vif superior to 10 were deleted. co-occurrence network analysis was performed on all sites for otus and phylum-class clusters following ju feng’s r and python scripts (ju et al., 2014; ju and zhang, 2015; hu et al., 2017) found at https://github.com/ richieju520/co-occurrence_network_analysis. only positive interactions between community members were considered. the spearman correlation and p-value cutoffs were set to 0.9 and 0.01 in the script. as for the c-score of the network, it was performed with the r package ecosimr (gotelli et al., 2015) based on a presence/absence otu matrix. the visualization and customization of the network was done with gephi software (bastian et al., 2009). finally, we used the computational method called pangenome-based functional profiles (panfp) (jun et al., 2015) to infer functional profiles for bacterial and archaeal classes. the software can be found at (https://github.com/srjun/panfp). generated kos where then analyzed using kegg website (https://www. genome.jp/kegg/ko.html) (kanehisa et al., 2016). pairwise intersection of kos between archaea and bacteria were obtained from the website molbiotools (http://www. molbiotools.com/listcompare.html). at last, the total shared kos were calculated using the dedicated bioinformatics and evolutionary genomics website tool, available at http://bioinformatics.psb.ugent.be/webtools/venn/. sequence data paired end fastq files, tags list, unfiltered bacterial and archaeal otu tables can be found at zenodo repository website following this link: http://doi.org/10.5281/zenodo.3678385 results the sequencing results along with the details regarding archaeal and bacterial reads and otus can be found in the supplementary material 1 and supplementary figures s2-s8. archaeal diversity and distribution the archaeal reads and otus represented 9% and 4% of the dataset, respectively. otu 2, assigned to nitrososphaeria (thaumarchaeota), constituted 97.4% of archaeal reads and was the second otu to hold the highest number of reads among bacterial and archaeal otus. overall, 5 phyla and 7 classes were detected after taxonomic assignment. furthermore, thaumarchaeota was dominant in all samples (supplementary fig. s4) followed by nanoarchaeaeota, mostly present in all samples and characterized by a lower number of reads. archaeal richness illustrated with the chao1 index (supplementary fig. s5) was similar to bacteria, and the highest values for this index were obtained at 200 m depth (p=1.859–05). evenness (pielou index) was significantly different between seasons (p=0.01758), especially between winter (february) and summer (august), but not between depths (supplementary table s1). archaeal diversity calculated with shannon, simpson and invsimpson indices (fig. 1a and supplementary table s1) also revealed significant differences between depths (p=0.006761), especially between near surface and deep waters (p=0.0028). similarly, depth was a factor explaining the difference in otus composition among samples as showed by the nmds (fig. 2a), anosim (p=0.001) and adonis (p=0.003) (supplementary table s2). undoubtedly, one otu contributed significantly to the beta diversity between treatments, i.e. otu 2 (thaumarchaeota, nitrososphaeria), in particular regarding depth (p=0.006761) since it drove respectively more than 45%, 47% and 40% of the difference between 2 and 15 m, 2 and 200 m and 15 and 200 m for the entire community (supplementary table s3). finally, the cca (fig. 3) performed on shl2 otus revealed that these 77 otus were positively correlated to total phosphorus, especially at 200 m, and also to conductivity, total nitrogen and ammoniacal nitrogen. in contrast, archaeal otus were negatively correlated to chlorophyll a, dissolved oxygen and total organic carbon. bacterial diversity and distribution forty-one bacterial phyla and 110 classes were assigned according to the pipeline. alphaproteobacteria (proteobacteria) yielded the highest number of otus (i.e. 411) (supplementary fig. s6) and actinobacteria the highest number of reads, i.e. 8,948,606 (supplementary fig. s7). actinobacteria, bacteroidetes, chloroflexi, cyanobacteria, planctomycetes and proteobacteria were highly present in all samples (fig. 4). samples at 200 m contained a higher phylum diversity than any other sample. indeed, when looking at the richness and diversity indices performed on the otus, the results showed a significant difference according to depth, while no significant diversity shifts were observed for sites or seasons (supplementary table s4). samples at 200 m displayed higher and significant values for all indices (chao1 pvalue=7.710–10, pielou p-value=4.706–6, shannon pvalue=4.322–08, simpson p-value=4.757–08 and invsimpson p-value=2.335–05) compared to the two other depths (2 and 15 m) (fig. 1-b and supplementary fig. s8). all investigated sites were not significantly different in otus composition (fig. 2-b) which was affected mainly by depth (anosim and adonis p-value were respectively 0.001 and 0.001) and by season (anosim and adonis p-value were 0.006 and 0.011, respectively) (suplake microbial statistical interactions 49 plementary table s5). the simper test (supplementary table s6) indicated that otus 4 and 5, assigned to the class anaerolineae (chloroflexi) and oxyphotobacteria (cyanobacteria) respectively, drove significantly the beta diversity. indeed, otu 4 contributed (p=1.714–05) to the disparity between surface and depth (up to 11%), as well as between 15 and 200 m depth (up to 12%). in addition, the average dissimilarity between 2 and 200 m was 65% and it was about 62% between 15 m and 200 m. otu 5 contributed significantly (p=0.0003058) to the disparity between february and the 3 other months (june 14%, august 14% and november 15%). the cca (fig. 5) that was performed on 1964 otus found at pt2, revealed that the variance is drawn in all directions. on one hand, total phosphorus revealed to be well linked to the otus present at 200 m. on the other hand, total nitrogen, chlorophyll a, conductivity and dissolved oxygen were linked to bacterial otus present at 2 and 15 m. relations between archaea and bacteria to assess possible interactions between bacteria and archaea we used two different and complementary approaches performed on the otus tables: a co-occurrence network analysis to determine significant connections between microbes and a functional profile prediction analysis to go further. the co-occurrence network (made of 260 and 24 nodes and 1230 and 33 undirected edges, for bacteria and archaea respectively) revealed connections among and between bacterial and archaeal otus. we only reported here strong (correlations ≥0.9) and significant (p<0.01) links. the c-score test with sim9 algorithm confirmed that the network was non-random. indeed, the observed c-score (10.523) was higher than the mean value (c-score mean=10.321, p<0.001) expected under the null model. on one hand, the archaeal otu 2 assigned to the genus nitrosopumilus (thaumarchaeota, nitrososphaeria, nitrosopumilales, nitrosopumilaceae) was linked to 8 bacterial otus. these otus were identified as anaerolineaceae (chloroflexi, anaerolineae, anaerolineales, otu 4), bsv26 (bacteroidetes, ignavibacteria, kryptoniales, otu 10), cl500-3 (planctomycetes, phycisphaerae, phycisphaerales, phycisphaeraceae, otu 19), nitrospira (nitrospirae, nitrospira, nitrospirales, nitrospiraceae, otu 25), imcc26256 (actinobacteria, acidimicrobiia, otu 46), sar11 clade (proteobacteria, alphaproteobacteria, otu 124), elsterales (proteobacteria, alphaproteobacteria, otu 135), and methylomonaceae (proteobacteria, gammaproteobacteria, methylococcales, otu 149). on the other hand, otu 67 identified as the archaeal class woesearchaeia (nanoarchaeaeota) was linked to 4 bacterial otus (fig. 6). these fig. 1. boxplot diagram comparing alpha diversity indices for archaea (a) and bacteria (b) (only significant results for indices are shown). for archaea, significant differences (p<0.01) are detected among the depth and season (tl1: february; tl2: june; tl3: august; tl4: november) variables. for bacteria, significant differences (p<0.01) are detected among depths. n, number of samples; a, b, significant differences. j.a. ezzedine et al.50 otus are identified as bsv26 (bacteroidetes, ignavibacteria, kryptoniales, otu 10), peribacteria (patescibacteria, gracilibacteria, otu 188), p2-11e (chloroflexi, otu 417) and kd4-96 (chloroflexi, otu 587). at the class level (supplementary fig. s9), nitrososphaeria (thaumarchaeota) was only linked to anaerolinae (chloroflexi). the woesearchaeia (nanoarchaeaeota) class maintained 3 links which were p2-11e (chloroflexi), kd4-96 (chloroflexi) and gracilibacteria (patescibacteria) and a new link with jg30.kf.cm66 (chloroflexi) appeared. the c-score test also confirmed that this network is non-random with observed c-score (5.1653) being higher than the mean value (c-score mean=4.4757, p<0.001) expected under the null model. the prediction of functional profiles from lineages (phylum – class) that are involved in the co-occurrence network were only possible for thaumarchaeota nitrososphaeria and its associated bacteria i.e. chloroflexi – anaerolineae; bacteroidetes – ignavibacteria; planctomycetes – phycisphaerae; nitrospirae – nitrospira; actinobacteria – acidimicrobiia; proteobacteria – alphaproteobacteria; proteobacteria – gammaproteobacteria (supplementary fig. s10). no functional profiles were found by the software for nanoarchaeaeota – woesearchaeia. nitrososphaeria had the minimum number of functional orthologs (ko=736), and gammaproteobacteria the highest (i.e. 5217). overall, the archaeal class nitrososphaeria had less kos involved in metabolic pathways, enzymes, carbon, sulfur and pyruvate metabolism, but methane and nitrogen metabolism. moreover, we calculated the shared number of kos among all these class and found a value of 186. when looking at the pairwise intersections (supplementary fig. s11) nitrososphaeria had the less shared kos with alpha and gammaproteobacteria (jaccard index=0.1231 and 0.0989, respectively), while most kos were shared among alpha and gammaproteobacteria (jaccard index=0.6833). this result showed that some pathways may be lacking in nitrososphaeria. discussion the choices we made shaped our observations we investigated the genetic diversity of bacteria and archaea at four months representing the different seasons of the year 2014. we are aware that the time scale of this study was relatively short (i.e. four dates) compared to other studies (bomberg et al., 2008; hugoni et al., 2013; parada and fuhrman, 2017) and the relatively low number of sampled sites, depths and months might not be representative to illustrate all the complex diversity of both archaea and bacteria in lake geneva. using illumina hiseq, although being very useful to assess deeply the 16s rrna-based genetic diversity of these prokaryotes, an important issue could be the primer specificity since we noted that the pair of primers targeting the v4-v5 region used in our study preferably amplified actinobacteria and thaumarchaeota. the in silico fig. 2. nmds plot illustrating separation of samples based upon differences in archaeal (a; stress value 0.03) and bacterial (b; stress value 0.05) otus structure. shape and color correspond respectively to different seasons february (tl1), june (tl2), august (tl3) and november (tl4), and depths (2, 15 and 200 m). for archaea, ellipses indicate 95% confidence intervals of otus grouped by different depths. for bacteria, the ellipses could not be generated. lake microbial statistical interactions 51 results showed that these primers had a good taxonomic resolution, covering 80.2% and 98.8% of the taxa at 0 and 3 mismatches respectively. it is noteworthy, however, that the behavior of primer pairs in vitro vs in silico may greatly differ as demonstrated in a previous study (ezzedine et al., 2020). ideally, different universal primers should be tested with the same sequencing technology to compare the results. it is also noteworthy that the taxonomic assignment is the result of bioinformatics algorithms, a procedure not perfect that can result in a variety of bias. also, the sequence identity threshold for 16s rrna otus can induce misidentification. as stated in the literature, sequences with 95% identity are likely to represent a same genus, whereas sequences with 97% identity may represent a same species (schloss and handelsman, 2005; schloss et al., 2009). in this study, we used otus, set the threshold at 96.6% and chose to analyze the taxonomic assignment to the genus level, whenever possible. indeed, the swarm algorithm was then used to cluster the sequences. the “d” parameter (i.e. the number of different nucleotides between sequences) was changed from 1 to 13 in order to be close to the identity threshold of 97% which describes same species. for the calculation, we took into account the median length of the sequences that was 377 bp. in addition, we chose stringent cutoff to filter the otus tables and considered only shared otus among pcr replicates. as we explored relationships between dominant archaea otus and other bacterial otus via a co-occurrence network analysis, we have to remind here again that it was constructed using a stringent cutoff (>0.9) in order to only reveal significant and strong correlations. in addition, we made the choice by focusing on positive correlation. the co-occurrence network along with the detected functional profiles from 16s rrna (and not from metagenome) are presented as a suggestion based on strong statistical analysis. regardless, this study sheds light on both bacterial and archaeal diversity in lake geneva using a high throughput sequencing approach and highlights, for this large and deep lake, possible relationships between prokaryotic communities, known to play a variety of key roles in aquatic ecosystems. archaeal and bacterial diversity our data set was mainly represented by two archaeal phyla, thaumarchaeota and nanoarchaeaeota, while arfig. 3. cca illustrating the separation of archaeal otus and samples based upon physicochemical descriptors. the cca was only performed on pt4 site were physicochemical variables were obtained. 12 samples and 77 archaeal otus were used in this analysis. j.a. ezzedine et al.52 chaea are generally more diversified in aquatic ecosystems (kozubal et al., 2013; evans et al., 2015; vanwonterghem et al., 2016; eme et al., 2017). other phyla were in fact detected, such as euryarchaeota, diapherotrites and crenarchaeota but they were clearly not dominant and only found with a low number of reads and otus. also, the diversity was found to be higher in deep-water (i.e. 200 m). globally, it has been reported that archaea may represent less than 10% of the microbial population in freshwater habitats (bomberg et al., 2008; bahram et al., 2019). our results are in accordance with these previous studies since archaea reads represented only 9% of the whole sequencing data set. this result could be related to the archaeal genome which is compact with a low number of ribosomal copies (koonin and wolf, 2008; angly et al., 2014), that may lead to underestimation of archaeal abundance (wurzbacher et al., 2017). in addition, the quite generic primers we used could have played a part in overlooking the archaeal diversity. as a matter of fact, (bahram et al., 2019) proved that employing an adequate pair of specific primers enabled to get accurate estimate of archaeal diversity by detecting readily, the asgard, the tack and the dpann superphyla in a variety of ecosystems. that being said, (haller et al., 2011) in their study of bacterial and archaeal communities at different contamination levels, using a cloning-sequencing approach also in lake geneva, reported that euryarchaeota were mainly found in contaminated sediments, rich in organic matter. (eme et al., 2017) found that ammonia oxidizing archaea (aoa) thaumarchaea living in the neighboring lake bourget, another (french) peri-alpine lake, were favored when ammonia concentrations were the lowest and fig. 4. stacked histogram of bacterial phylum. phylum abundances are represented by read relative abundance per sample. phyla with less than 5% of read relative abundance in a sample are merged and identified as “others”. lake microbial statistical interactions 53 during winter. for the same lake, other studies reported that archaea can occupy different ecological niches due to the influence of different abiotic and biotic factors. indeed, archaeal otus were shown to be structured by factors such as ammoniacal nitrogen, total nitrogen and total phosphorus (berdjeb et al., 2013; pollet et al., 2018). at last, (parada and fuhrman, 2017) reported that depth and seasonality influenced archaeal community composition, and that archaeal group selected different ecotypes depending on how they fit to environmental conditions at each depth. overall, this suggests that archaea have preferential niches and that thaumarchaea of lake geneva are most probably adapted to thrive at the three sampled depths and the studied months, compared to other groups. for instance, otu 2 that contributed to sample dissimilarity also held the highest number of reads for archaea and the second in the data set. this otu was assigned to candidatus nitrosopumilus from the family of the nitrosopumilaceae (thaumarchaeota), a genus reported to be ubiquitous in marine environments and an important ammonia oxidizing archaea. bacteria were much more diverse than archaea. actinobacteria, cyanobacteria, proteobacteria, chloroflexi and bacteroidetes dominated this community. this result agrees with previous studies which identified most of these groups as predominant phyla of freshwater systems (glöckner et al., 1999; debroas et al., 2009; newton et al., 2011; zwirglmaier et al., 2015). as for the archaea, depth mainly explained the differences found in the bacterial diversity and the highest values in bacteria richness and diversity were also found at 200 m depth (with a dominance of actinobacteria). no difference in richness, diversity and evenness was found among the different sampling sites. however, the season also played a role since bacterial otus found at 2 and 15 m (but not those at 200 m) were segregated by months. otu 5 (oxyphotobacteria) was the most influential otu to drive the beta diversity within seasons. otu 4 (anaerolineae) was more present at 200 m and it drove the beta diversity between 2 and 200 m, and 15 and 200 m. physicochemical descriptors explained 41% of the bacterial otus distribution present at pt4. total phosphorus explained well the presence of some otus at 200 m, a result reminding the study of (berdjeb et al., 2011) who reported that the bacterial community structure present in the hypoand epilimion in two other peri-alpine lakes, bourget and annecy, was affected mainly by bottom-up factors regardless of the local environmental conditions of these lakes. fig. 5. cca illustrating the separation of bacterial otus and samples based upon physicochemical variables. the cca was only performed on pt4 site were physicochemical variables were obtained. 12 samples and 1,964 bacterial otus were used in this analysis. j.a. ezzedine et al.54 potential relationships between prokaryotes a variety of relationships was found between the prokaryotes, suggesting potential ecological and/or biogeochemical interactions. the archaeal otu 2 assigned to the phylum – class; thaumarchaeota – nitrososphaeria co-occurred with 8 bacterial otus. recently, in the release of silva ssu database 138 (december 16, 2019), nitrososphaeria was transferred to the crenarchaeota phylum. however, the analysis was performed with database 132 and prior to the release of the new database version, explaining why we maintained the former classification. the detection of functional profiles revealed that nitrososphaeria is less equipped in term of pathway in comparison to associate bacteria. this is expected since archaea possesses a compact genome. in addition, only few kos were common between nitrososphaeria and the other bacteria, suggesting that some functions are missing in the archaeal genome. these results explained probably why nitrososphaeria co-occurred with other bacteria. the archaeal otu 2 (i.e. nitrosopumilus), mentioned above, cooccurred with the bacterial otu 25 assigned to nitrospira (nitrospirae). this relationship is coherent with the finding of (parada and fuhrman, 2017) who also reported a correlation between marine thaumarchaea and ammonia oxidizing bacteria (aob) such as nitrospina, both responsible for nitrification. archaeal otu 2 displayed also a link with the bacterial otu 19, identified as cl500-3 (i.e. a planctomycetes). these bacteria were already reported to be associated to aoa such as thaumarchaeota in oxygenated hypolimnion of deep freshwater lakes (okazaki et al., 2017). nitrosopumilus is known to operate the methane cycle and produce methylphosphonic acid (carini et al., 2014). methane-oxidizing archaea (moa) and bacteria (mob) such as αand γ-proteobacteria (hanson and hanson, 1996; lüke et al., 2010) can either produce or consume methane. these microorganisms are essential for the methane release and rate in the atmosfig. 6. co-occurrence network between bacteria and archaea otus. only otus that passed the cutoff (correlation coefficient ≥0.9 and p<0.01) are illustrated. only positive relationships are shown. red and salmon color are the connections from the archaeal otu 2 to its associate bacteria. while blue color represents the connections between archaeal otu 67 and its associate bacteria. the purple color indicates that both archaeal otu shared a connection with a bacterial otu. lake microbial statistical interactions 55 phere. therefore, the occurrence between the archaeal otu 2 and the bacterial otu 4 assigned to anaerolineaceae (chloroflexi) makes sense, with the possibility that anaerolineaceae provide organic acid such as acetate to other microorganisms like acetoclastic methanogens. moreover, anaerolineaceae could form a syntrophic cooperation with some archaea involved in methanogenic degradation of alkanes (lüke et al., 2010; liang et al., 2015). similarly, we observed mob co-occurrence through otu 149 assigned to the methylomonaceae (gammaproteobacteria). furthermore, a mutualistic interaction might be occurring between archaea otu 2 and bacteria otu 124 assigned to sar11 clade (alphaproteobacteria). (parada and fuhrman, 2017) a link between some thaumarchaea and sar groups such as sar406 and sar86 was also found in the ocean. in fact, nitrosopumilus produce methylphosphonic acid (mpn), which is decomposed by phosphate-starved microorganisms such as sar11 clade (chemoheterotrophic bacteria) that produce ch4 from mpn when pi (inorganic phosphate) starved (carini et al., 2014). this was also suggested by the functional profile where nitrososphaeria, alpha and gammaproteobacteria shared 526 to 536 from 736 kos pointing out that the bacterial class lack some pathways that can be compensated with an interaction with nitrososphaeria. finally, the co-occurrence with otu 10 assigned to bsv26 (kryptoniales) and otu 46 identified as imcc26256 (acidimicrobiia) could be ecologically relevant since these two otus were originally isolated together from extremely acidic environments or geothermal sites were archaea blossom (ludwig et al., 2012; hu et al., 2017). the second most important archaeal otu (i.e. otu 67) showed important relationships with bacterial otus 10, 188, 417 and 587. otu 67 was assigned to the class woesearchaeia of the phylum nanoarchaeaeota. woesearchaeota are widely spread in diverse environments and (liu et al., 2018) reported a syntrophic relationship between woesearchaeota and other methanogenic archaea. this syntrophic and/or mutualistic partnership emerged from deficiencies in metabolic pathways within the woesearchaeota class. (liu et al., 2018)also suspected that woesearchaeota develop a partnership with some bacterial methanogens, proposing a woesearchaeotamethanogens consortium, likely to influence methane formation. however, these authors lacked reliable bacterial data to analyze collectively these microorganisms and prove this assumption. our results support their finding. a relationship also existed with chloroflexi otu 417 (p2-11e), 587 (kd4-96) and to jg30.kf.cm66 (supplementary fig. s9). in fact, chloroflexi possess a wide diversity of metabolisms (e.g. aerobic respiration, denitrification and phototrophy), and ecological roles, but are best known as photoheterotrophs (ward et al., 2018). they are reported to be dominant and abundant in methanogenic reactors (bovio et al., 2019). recently, (ward et al., 2019) reported that within the chloroflexi phylum some of them such as ohk40 are capable of methane oxidation (photomethanotrophy). we cannot pinpoint the exact need or relationship between otu 67 and the other bacterial otus but a syntrophic or/and a mutualistic link is very probable between them. this could explain the abundance of chloroflexi observed at 200 m (fig. 1) where oxygen concertation was found to be lower than at the surface (i.e. 7 vs 9-10 mg l–1). indeed, if some chloroflexi otus can produce methane they are likely to occur in less oxic layers since methanogenesis is not favored where oxygen is abundant (carini et al., 2014), and, as pointed out by (mayr et al., 2019), mob can occupy different niches according to oxygen and methane concentration. at last, the co-occurrence observed with otu 188 assigned to peribacteria (gracilibacteria) also make sense since the latter lack some pathways and is predicted to be either a symbiont or closely dependent on other community members for key building blocks (sieber et al., 2019). conclusions in parallel to studies dealing with physiology, biodiversity, biological indication and assessment, assessing interactions between microorganisms in pelagic communities are of great ecological interest but still remain poorly documented in lakes. using illumina hiseq combined with 16s metabarcoding via universal primer, cooccurrence networks and functional profile predictions, we showed that lake geneva is booming with connections, and our results suggest that some otus co-occur for scavenging or sharing similar ecological niches for mutual benefits. sequencing different functional genes and conducting experiments with isolated organisms could help to better understand the relationship between key “species” and biotic interactions sustaining the microbial ecosystem functioning. acknowledgments we are grateful to the editor and the reviewer for helping us to improve the article. we would like to thank pascal perney, philippe quetin, and jean-christophe hustache for sampling. we are also thankful to louis jacas, cécile chardon, eric capo and gaétan maechler for technical assistance during the project. we thank julie gueguen, valentin vasselon, viet tran-khac and françois kek for bioinformatics and statistical help. finally, special thanks are attributed to dr. se-ran jun for providing some help to generate kos from the otus tables. j.a. ezzedine et al.56 corresponding author: stephan.jacquet@inrae.fr keywords: bacteria; archaea; lake geneva; co-occurrence; diversity; functional profiles. authors contributions: sj, designed the study and collected the samples; je, performed bioinformatics and statistical analysis; je, sj, analyzed and interpreted the results; je, sj wrote the manuscript with revisions by yd. conflict of interest: the authors declare that the research was conducted in the absence of any commercial or financial relationships that could be construed as a potential conflict of interest. received: 12 may 2020. accepted: 18 august 2020. this work is licensed under a creative commons attribution noncommercial 4.0 license (cc by-nc 4.0). ©copyright: the author(s), 2020 licensee pagepress, italy advances in oceanography and limnology, 2020; 11:9099 doi: 10.4081/aiol.2020.9099 references alonso-sáez l, waller as, mende dr, bakker k, farnelid h, yager pl, lovejoy c, tremblay jé, potvin m, heinrich f, estrada m, riemnann l, bork p, pedros-alio c, bertilsson s, 2012. role for urea in nitrification by polar marine archaea. proc. natl. acad. sci. u. s. a. 109: 17989– 17994. angly fe, dennis pg, skarshewski a, vanwonterghem i, hugenholtz p, tyson gw, 2014. copyrighter: a rapid tool for improving the accuracy of microbial community profiles through lineage-specific gene copy number correction. microbiome 2:1–13. anneville o, souissi s, ibanez f, ginot v, druart jc, angeli n, 2002. temporal mapping of phytoplankton assemblages in lake geneva: annual and interannual changes in their patterns of succession. limnol. oceanogr. 47: 1355–1366. bahram m, anslan s, hildebrand f, bork p, tedersoo l, 2019. newly designed 16s rrna metabarcoding primers amplify diverse and novel archaeal taxa from the environment. environ. microbiol. rep. 11: 487–494. bastian m, heymann s, jacomy m, 2009. gephi: an open source software for exploring and manipulating networks. bt international aaai conference on weblogs and social. int. aaai conf. weblogs soc. media: 361–362. berdjeb l, ghiglione jf, jacquet s, 2011. bottom-up versus topdown control of hypo-and epilimnion free-living bacterial community structures in two neighboring freshwater lakes. appl. environ. microbiol. 77: 3591–3599. berdjeb l, pollet t, chardon c, jacquet s, 2013. spatio-temporal changes in the structure of archaeal communities in two deep freshwater lakes. fems microbiol. ecol. 86: 215–230. bomberg m, montonen l, münster u, jurgens g, 2008. diversity and function of archaea in freshwater habitats. curr. trends microbiol. 4: 61–89. bovio p, cabezas a, etchebehere c, 2019. preliminary analysis of chloroflexi populations in full-scale uasb methanogenic reactors. j. appl. microbiol. 126: 667–683. braga rm, dourado mn, araújo wl, 2016. microbial interactions: ecology in a molecular perspective. brazilian j. microbiol. 47: 86–98. carini p, white ae, campbell eo, giovannoni sj, 2014. methane production by phosphate-starved sar11 chemoheterotrophic marine bacteria. nat. commun. 5: 1–7. casamayor eo, 2017. towards a microbial conservation perspective in high mountain lakes, p. 157–180. in: catalan j, ninot jm and aniz mm (eds). high mountain conservation in a changing world. springer, cham. cavicchioli r, 2011. archaea timeline of the third domain. nat. rev. microbiol. 9: 51–61. debroas d, humbert jf, enault f, bronner g, faubladier m, cornillot e, 2009. metagenomic approach studying the taxonomic and functional diversity of the bacterial community in a mesotrophic lake (lac du bourget france). environ. microbiol. 11: 2412–2424. dekas ae, poretsky rs, orphan vj, 2009. deep-sea archaea fix and share nitrogen in methane-consuming microbial consortia. sci. mag. 326: 422–427. dinno a, 2017. dunn’s test of multiple comparisons using rank sums. eme l, spang a, lombard j, stairs cw, ettema tjg, 2017. archaea and the origin of eukaryotes. nat. rev. microbiol. 15: 711–723. evans pn, parks dh, chadwick gl, robbins sj, orphan vj, golding sd, tyson gw, 2015. methane metabolism in the archaeal phylum bathyarchaeota revealed by genome-centric metagenomics. science (80-. ). 350: 434–438. ezzedine ja, chardon c, jacquet s, 2020. new 16s rrna primers to uncover bdellovibrio and like organisms diversity and abundance. j. microbiol. methods 175: 105996. fuhrman ja, davis aa, 1997. widespread archaea and novel bacteria from the deep sea as shown by 16s rrna gene sequences. mar. ecol. prog. ser. 150: 275–285. glöckner fo, fuchs bm, amann r, 1999. bacterioplankton compositions of lakes and oceans: a first comparison based on fluorescence in situ hybridization. appl. environ. microbiol. 65: 3721–3726. goral f, schellenberg j, 2017. goeveg: functions for community data and ordinations. gotelli n, hart e, ellison a, 2015. ecosimr: null model analysis for ecological data. haller l, tonolla m, zopfi j, peduzzi r, wildi w, poté j, 2011. composition of bacterial and archaeal communities in freshwater sediments with different contamination levels (lake geneva, switzerland). water res. 45: 1213–1228. hanson rs, hanson te, 1996. methanotrophic bacteria. microbiol. rev. 60: 439–471. hu a, ju f, hou l, li j, yang x, wang h, mulla si, sun q, burgmann h, yu c-p, 2017. strong impact of anthropogenic contamination on the co-occurrence patterns of a riverine microbial community. environ. microbiol. 19: 4993–5009. hu d, cha g, gao b, 2018. a phylogenomic and molecular markers based analysis of the class acidimicrobiia. front. microbiol. 9: 1–12. hugoni m, etien s, bourges a, lepère c, domaizon i, mallet lake microbial statistical interactions 57 c, bronner g, debroas d, mary i, 2013. dynamics of ammonia-oxidizing archaea and bacteria in contrasted freshwater ecosystems. res. microbiol. 164: 360–370. jacquet s, domaizon i, anneville o, 2014. the need for ecological monitoring of freshwaters in a changing world: a case study of lakes annecy, bourget, and geneva. environ. monit. assess. 186: 3455–3476. ju f, zhang t, 2015. bacterial assembly and temporal dynamics in activated sludge of a full-scale municipal wastewater treatment plant. isme j. 9: 683–695. ju f, xia y, guo f, wang z, zhang t, 2014. taxonomic relatedness shapes bacterial assembly in activated sludge of globally distributed wastewater treatment plants. environ. microbiol. 16: 2421–2432. jun sr, robeson ms, hauser lj, schadt cw, gorin aa, 2015. panfp: pangenome-based functional profiles for microbial communities. bmc res. notes 8: 1–7. kanehisa m, sato y, kawashima m, furumichi m, tanabe m, 2016. kegg as a reference resource for gene and protein annotation. nucleic acids res. 44: d457–d462. koizumi y, takii s, fukui m, 2004. depth-related change in archaeal community structure in a freshwater lake sediment as determined with denaturing gradient gel electrophoresis of amplified 16s rrna genes and reversely transcribed rrna fragments. fems microbiol. ecol. 48: 285–292. könneke m, bernhard ae, de la torre jr, walker cb, waterbury jb, stahl da, 2005. isolation of an autotrophic ammonia-oxidizing marine archaeon. nature 437: 543–546. koonin e v., wolf yi, 2008. genomics of bacteria and archaea: the emerging dynamic view of the prokaryotic world. nucleic acids res. 36: 6688–6719. kozubal ma, romine m, jennings rd, jay zj, tringe sg, rusch db, beam jp, mccue la, inskeep wp, 2013. geoarchaeota: a new candidate phylum in the archaea from hightemperature acidic iron mats in yellowstone national park. isme j. 7: 622–634. liang b, wang ly, mbadinga sm, liu jf, yang sz, gu jd, mu bz, 2015. anaerolineaceae and methanosaeta turned to be the dominant microorganisms in alkanes-dependent methanogenic culture after long-term of incubation. amb express 5:37. linz a, 2018. otutable: north temperate lakes microbial observatory 16s time series data and functions. available from: https://rdrr.io/cran/otutable/ liu x, li m, castelle cj, probst aj, zhou z, pan j, liu y, banfield jf, gu jd, 2018. insights into the ecology, evolution, and metabolism of the widespread woesearchaeotal lineages. microbiome 6: 1–16. ludwig w, euzéby j, schumann p, busse h-j, trujillo me, kämpfer p, whitman wb, 2012. road map of the phylum actinobacteria. bergey’s manual® syst. bacteriol.: 1–28. lüke c, krause s, cavigiolo s, greppi d, lupotto e, frenzel p, 2010. biogeography of wetland rice methanotrophs. environ. microbiol. 12: 862–872. mahé f, rognes t, quince c, de vargas c, dunthorn m, 2015. swarmv2: highly-scalable and high-resolution amplicon clustering. peerj 2015: 1–12. martin m, 2011. cutadapt removes adapter sequences from highthroughput sequencing reads. embnet.journal 17: 1–3. mayr mj, zimmermann m, guggenheim c, brand a, bürgmann h, 2019. niche partitioning of methane-oxidizing bacteria along the oxygen–methane counter gradient of stratified lakes. isme j.: 274–287. newton rj, jones se, eiler a, mcmahon kd, bertilsson s, 2011. a guide to the natural history of freshwater lake bacteria. microbiol. mol. biol. rev. 75:14–49. okazaki y, fujinaga s, tanaka a, kohzu a, oyagi h, nakano si, 2017. ubiquity and quantitative significance of bacterioplankton lineages inhabiting the oxygenated hypolimnion of deep freshwater lakes. isme j. 11: 2279–2293. oksanen j, blanchet fg, friendly m, kindt r, legendre p, mcglinn d, minchin pr, o’hara rb, simpson gl, solymos p, stevens mhh, szoecs e, wagner h, 2019. package ‘vegan’ title community ecology package. community ecol. packag. 2: 1–297. ortiz-alvarez r, casamayor eo, 2016. high occurrence of pacearchaeota and woesearchaeota (archaea superphylum dpann) in the surface waters of oligotrophic high-altitude lakes. environ. microbiol. rep. 8: 210–217. pacheco ar, segrè d, 2019. a multidimensional perspective on microbial interactions. fems microbiol. lett. 366: 1–11. parada ae, fuhrman ja, 2017. marine archaeal dynamics and interactions with the microbial community over 5 years from surface to seafloor. isme j. 11: 2510–2525. pollet t, berdjeb l, chardon c, jacquet s, 2018. contrasting temporal patterns in ammonia-oxidizing archaeal community dynamics in two peri-alpine lakes with different trophic status. aquat. microb. ecol. 81: 95–108. quast c, pruesse e, yilmaz p, gerken j, schweer t, glo fo, yarza p, 2013. the silva ribosomal rna gene database project : improved data processing and web-based tools. nucleic acids res. 41: 590–596. r core team, 2019. r: a language and environment for statistical computing (v. 3.5.0). r foundation for statistical computing, wien, austria. robertson ce, harris jk, spear jr, pace nr, 2005. phylogenetic diversity and ecology of environmental archaea. curr. opin. microbiol. 8: 638–642. rognes t, flouri t, nichols b, quince c, mahé f, 2016. vsearch: a versatile open source tool for metagenomics. peerj 4: 1–22. schloss pd, handelsman j, 2005. introducing dotur, a computer program for defining operational taxonomic units and estimating species richness. appl. environ. microbiol. 71:1501–1506. schloss pd, westcott sl, ryabin t, hall jr, hartmann m, hollister eb, lesniewski ra, oakley bb, parks dh, robinson cj, sahl jw, stres b, thallinger gg, van horn dj, weber cf, 2009. introducing mothur: open-source, platform-independent, community-supported software for describing and comparing microbial communities. appl. environ. microbiol. 75: 7537–7541. seyler lm, tuorto s, mcguinness lr, gong d, kerkhof lj, 2019. bacterial and archaeal specific-predation in the north atlantic basin. front. mar. sci. 6: 1–10. sieber cmk, paul bg, castelle cj, hu p, tringe sg, valentine dl, anderson gl, banfield jf, 2019. unusual metabolism and hypervariation in the genome of a gracilibacteria (bd15) from an oil degrading community. biorxiv. doi: 10.1101/595074. j.a. ezzedine et al.58 torondel b, ensink jhj, gundogdu o, ijaz uz, parkhill j, abdelahi f, nguyen va, sudgen s, gibson w, walker aw, quince c, 2016. assessment of the influence of intrinsic environmental and geographical factors on the bacterial ecology of pit latrines. microb. biotechnol. 9:209–223. vanwonterghem i, evans pn, parks dh, jensen pd, woodcroft bj, hugenholtz p, tyson gw, 2016. methylotrophic methanogenesis discovered in the archaeal phylum verstraetearchaeota. nat. microbiol. 1: 1–9. varela mm, van aken hm, sintes e, herndl gj, 2008. latitudinal trends of crenarchaeota and bacteria in the mesoand bathypelagic water masses of the eastern north atlantic. environ. microbiol. 10: 110–124. vuillemin a, wankel sd, coskun ök, magritsch t, vargas s, estes er, spivack aj, smith dc, pockalny r, murray rw, d’hondt s, orsi wd, 2019. archaea dominate oxic subseafloor communities over multimillion-year time scales. sci. adv. 5: 1–12. wang f, men x, zhang g, liang k, xin y, wang j, li a, zhang h, liu h, wu l, 2018. assessment of 16s rrna gene primers for studying bacterial community structure and function of aging flue-cured tobaccos. amb express 8: 1–9. wang y, qian py, 2009. conservative fragments in bacterial 16s rrna genes and primer design for 16s ribosomal dna amplicons in metagenomic studies. plos one 4:e7401. ward l, shih pm, hemp j, kakegawa t, fischer ww, mcglynn se, 2019. phototrophic methane oxidation in a member of the chloroflexi phylum. biorxiv: 531582. doi: 10.1101/ 531582. ward lm, hemp j, shih pm, mcglynn se, fischer ww, 2018. evolution of phototrophy in the chloroflexi phylum driven by horizontal gene transfer. front. microbiol. 9: 1–16. wickham h, chang w, henry l, pedersen tl, takahashi k, wilke c, woo k, 2018. ggplot2: create elegant data visualisations using the grammar of graphics. available from: https://rdrr.io/cran/ggplot2/ wurzbacher c, fuchs a, attermeyer k, frindte k, grossart hp, hupfer m, casper p, monaghan mt, 2017. shifts among eukaryota, bacteria, and archaea define the vertical organization of a lake sediment. microbiome 5: 1–16. zwirglmaier k, keiz k, engel m, geist j, raeder u, 2015. seasonal and spatial patterns of microbial diversity along a trophic gradient in the interconnected lakes of the osterseen lake district, bavaria. front. microbiol. 6: 1–18. layout 1 introduction as human beings we realize our living immersed in networks of conversations, which we are not always aware of, that represent both a treasure and a danger. it is actually through the voices that we are listening to, and through the narratives we embrace, that we learn and reflect about the natural world and that we perceive our deep and vital interconnection with or separation from it. we should recognize, both as scientists and as citizens, that we can chose the story and the conversation we take part, because this shapes our responses to any crisis and challenge (maturana and davila, 2009). in long-term ecological research (lter) we are engaged in conversations with multiple actors: organisms, ecosystems, methodologies, data, other scientists, citizens, and other typologies of stakeholders, are some of the “voices” that we receive, interpret, and express, creating our experience and knowledge. at the same time, these voices shape and are influenced by our “inner world” the phycological sphere defined by our emotions, thoughts, identities and beliefs – which is considered one of the main drivers of changes in our behaviour and values, enabling actions in uncertain or unprecedented situations (wamsler and brink, 2018; ives et al., 2018, 2019). indeed, science is not only an individual intellectual or mental process: it is a collective, both organizational and institutional, practice driven by social dynamics. in this pursuing practice, my and other authors’ perspectives indicate that the “inner world” is a necessary and positive player (barbalet, 2002; koppman et al., 2015). grounding on the last decade activities of the italian lter network (lter-italy; www.lteritalia.it), i will review and report here some experiences, mainly focussing on those “voices from the water” that have been relevant and helpful to delineate my perspective on the role that ecology (i.e., the branch of science which studies the interrelationships of organisms and their environments, in the complex interplay of functions and processes, information cycling, and cooperative and competitive dynamics) is challenged to play in a world of rapid change, characterized by socio-ecological conflicts and economic contradictions. the one from plankton (fig. 1) is the “voice” that i will primarily listen to, as that key ecological community of generally microscopic organisms floating in the waters of the earth (hensen, 1887; d’alelio, 2020) is the main object of my study along the years. an italian poet, pierluigi cappello (2013) writes: “there are words without a body and words with a body. freedom is a world without a body. like soul. like love. they need someone who lends them flesh, blood and boundaries to make them concrete”. i like to picture that this is what we do with plankton: we lend our bodies and voices to make them become visible and real, appearing from the waters. in particular, this paper describes and highlights article voices from the water: experience, knowledge, and emotions in long-term ecological research (lter italy) alessandra pugnetti cnr-ismar, arsenale tesa 104, castello 2737/f, 30122 venezia, italy abstract based on the last decade of activities at the italian long-term ecological research network (lter-italy), i describe and highlight here some major outcomes and challenges, by picturing different “voices”, which we are listening to and we are talking with. organisms, ecosystems, methodologies, data, researchers, stakeholders and citizens: their voices – i.e., the information which we receive from them, and then interpret and, consequently, express – create our experience and knowledge, which we share with and convey to our contemporaries and future generations. one of the main narrator’s voices will be that of plankton: how we “listen” to them, describe and share long-term data and researches, also with the wide public. through the “voices from the water” i will report and discuss experiences, which have been relevant also to open up the views on the role that science is challenged to play in a world of rapid change, characterized by complexity and contradictions. in particular i will consider: i) those voices coming from various lter aquatic sites, mainly addressing the comparison among them; ii) how to make these voices most harmonized and audible through the open science approach; and iii) how to put the lter voices in an effective dialogue with society. finally, i will share some thoughts about the necessity and the possibility to open the purely scientific cognitive approach to other forms of knowledge, related to our intuition, feelings and empathy. if you are a poet, you will see clearly that there is a cloud floating in this sheet of paper. without a cloud, there will be no rain; without rain, the trees cannot grow; and without trees, we cannot make paper. (…) if the cloud is not here, the sheet of paper cannot be here either. so we can say that the cloud and the paper inter-are. (…). “to be” is to inter-be. we cannot just be by ourselves alone. we have to inter-be with every other thing. (tich nath han) no nco mm er cia l u se on ly a. pugnetti60 some major outcomes and challenges, in the form of the different voices, which we are listening to and we are conveying within and beyond our community, i.e., the one of scientists. in pursuing this aim, i will be: i) listening to the voices coming from the lter aquatic sites, mainly addressing the comparison among them; ii) suggesting how to make these lter voices more harmonized and audible through the open science approach; and, finally, iii) illustrating my perspective on how getting the lter voices engaged in an effective dialogue with society. in the end, i will share some thoughts about the chance, and necessity, to open the purely scientific cognitive approach to other forms of knowledge, integrating our inner world. voices from lter-italy lter-italy aquatic sites and plankton the principal aim of lter is understanding, analysing, and monitoring changes in ecosystem patterns and processes over extended periods of time, typically decades. lter is prevalently organised in networks of sites and platforms, at the national (e.g., lter-italy), continental (e.g., european, lter-europe: http://www.ltereurope.net/) and global level (lter-international, ilter: www.ilter.network), where approaches and interpretations of ongoing ecological processes are developed (mirtl et al., 2018; mollenhauer et al., 2018), also with the aim of creating a legacy of well-designed and documented knowledge for future generations. amongst the many papers that have been issued in the last decades about lter research, networks and sites, i have been – and i still am – motivated by two of them, which are the voices providing me with both the knowledge and the language words and metaphors to convey it. those seminal papers and their authors are still influencing and nourishing my inner world as well, supporting my emotional perspective and my long-lasting bond with lter. the first paper was written by magnuson (1990) and gifted me with the metaphor of “the invisible present”, which express powerfully the role of lter in connecting the current environmental (both physical and biological) conditions to the past, in order to understand where we are and where we wish to move toward. in this paper, there is also a clear call for responsibility in lter: the invisible present is actually defined as the time scale within which our responsibilities for the earth are most evident since, within this time scale, ecosystems change during our lifetimes and the lifetimes of our children and our grandchildren. the second seminal paper was written by peters (2010) and made me understand and appreciate the value of lter sites, by bringing together their role along three different dimensions and ecological strategies: the long (long-term studies), the deep (short-term, patternprocess studies for deep understanding) and the broad (observation networks of sites for broad-scale patterns). the fig. 1. optic microscopy view of a mixed winter phytoplankton (left) and zooplankton (right) community gathered at the lter site gulf of venice, northern adriatic sea. pictures courtesy of stefania finotto (phytoplankton), and elisa camatti and marco pansera (zooplankton). no nco mm er cia l u se on ly voices from the water 61 process of construction and consolidation of the italian lter network, to which i contributed along with many other colleagues, have always been guided for me by the voices expressed by these two authors, and have provided me the foundation of any further conceptual enrichment. the lter-italy network has been established in 2006, year that marks its official affiliation to ilter (i.e., the ‘international’ lter network). the story of lter-italy has been driven by professional collaboration and friendship, with many events and memories that promoted the long-lasting connection within this wide community. in the past 15 years, the development of lter-italy has been tightly connected with that of lter-europe, even more now that the european network has entered a new phase, while developing the elter research infrastructure (elter ri) as a formally recognised esfri research infrastructure, since 2018. lter-italy currently consists of 79 research sites, organized in 25 parent sites (i.e., made by multiple research sites), managed and coordinated by public research institutions, universities, and environmental agencies. the sites represent the main ecosystem typologies of italy: they include terrestrial, freshwater, transitional and coastal marine environments, giving the network a marked interdisciplinary brand. forty sites, i.e. more than half of the total lter-italy sites, are aquatic (tab. 1). they are distributed along the whole peninsula and they include the most common italian lake typologies (large and deep subalpine lakes, small and shallow mountain lakes and reservoirs), the main italian lagoons and relevant marine coastal ecosystems. long-term series of data on plankton (phyto-and/or zooplankton) dynamics have been collected during the last forty years at most lter-italy aquatic sites, providing an invaluable empirical and rigorous knowledge for the sustainable management of aquatic resources, considering that in the last decades most aquatic ecosystems in italy have undergone significant ecological changes. the voices from lter-italy aquatic sites – from the large subalpine lakes to mountain and artificial lakes, from lagoons to marine coastal ecosystems – were gathered and compared by 54 authors in a first, and unique until now, synoptic and trans-domain overview of planktab. 1. list of the lter-italy aquatic sites involved in the papers cited in the text and/or in the cammini lter initiative. for the full list of the lter-italy aquatic sites see: www.lteritalia.it lter-italy lter-italy ecosystem morabito rogora zingone tanentzap pilotto cammini parent site research site typology et al., et al., et al., et al., et al., lter 2018 2018 2019 2020 2020 lake ecosystem of sardinia lake bidighinzu artificial lake x x lake ecosystem of sardinia lake sos-canales artificial lake x southern alpine lakes lake como deep subalpine lake x southern alpine lakes lake garda deep subalpine lake x x southern alpine lakes lake iseo deep subalpine lake x southern alpine lakes lake maggiore deep subalpine lake x x x southern alpine lakes lake orta deep subalpine lake x mountain lakes lago anterselva alpine lake x mountain lakes lago piccolo di monticoli alpine lake x mountain lakes lake paione superiore alpine lake x x mountain lakes lake paione inferiore alpine lake x x mountain lakes lake santo parmense appennine lake x x mountain lakes lake scuro parmense appennine lake x x mountain lakes lake tovel alpine lake x x lake trasimeno shallow lake x lagoons of salento alimini coastal lagoon x lagoons of salento aquatina coastal lagoon x marine ecosystems of sardinia lagoon of cabras coastal lagoon x mar piccolo of taranto coastal lagoon x x lagoon of venice coastal lagoon x x x gulf of naples marechiara mediterranean shelf and sea x x x marine ecosystems of sardinia gulf of olbia mediterranean shelf and sea x northern adriatic sea gulf of trieste mediterranean shelf and sea x x northern adriatic sea gulf of venice mediterranean shelf and sea x x x ligurian sea portofino promontory mediterranean shelf and sea x northern adriatic sea senigallia-susak transect mediterranean shelf and sea x no nco mm er cia l u se on ly a. pugnetti62 ton dynamics, based mainly on published studies, but also checked and complemented with unpublished information (morabito et al., 2018). this paper is indissolubly associated to the figure of giuseppe morabito, who, during the paper drafting, passed away for a sudden upsurge of the illness that hit him some years before. giuseppe was a strong and much appreciated voice in phytoplankton ecology, both in italy and abroad. but, above all, he was a friend for most of us. the “voice” of giuseppe, both the remembrance of its tone and laughs and the ecological conceptualization it brought to us, accompanied us for many months following his death and it was crystalized in our paper. in the article by morabito et al. (2018), a general conceptual framework was achieved, connecting the large and small scales in a chain of events embracing the whole ‘aquatic continuum’, where the plankton responses appear to depend mainly on the trophic state of the ecosystems and on the strength of the local anthropogenic disturbance. actually, the long-term patterns were linked to a variety of factors, differently identified in: i) the physical and meteorological conditions; ii) the combination of bottom-up and top-down controlling factors; iii) the progressive improvement of wastewater treatment; and iv) the contemporary decrease of precipitations and nutrient loading from inflowing rivers. besides, the marked increase in water temperature over the last decades was a trait shared by many italian marine and freshwater environments: these results concur to confirm the significant warming of marine ecosystems and largest lakes, in line with the outcome of measurements recorded at the global scale. some of the voices of lter-italy aquatic sites (tab. 1) joined the ones coming from the terrestrial sites, in two cross-domain studies conducted at the european level (rogora et al., 2018; pilotto et al., 2020). the first of these studies (rogora et al., 2018) offers a comprehensive overview and synthesis on the effects of climate change in different mountain ecosystems, considering several lter sites, both terrestrial and freshwater (lakes and streams), distributed along the apennines (italy) and the alps (italy, switzerland and austria). rogoraet al. (2018) focus on the change in vegetation cover and carbon uptake, on the alteration of biogeochemical cycles in soils and water, and on the change in phenology and biological diversity in mountain ecosystems; data showed varying levels of effects in response to climate change, which has a prevalent indirect impact, resulting from multiple interactions. for instance, the interannual variability in the duration of the snow cover appeared to play a key role, affecting nutrient cycles both in soils and in surface waters and influencing the biological communities as well. the second cross-site study mentioned herein (pilotto et al., 2020) is a meta-analysis of multidecadal biodiversity series in europe, aiming to fully understand the patterns of local biodiversity change and the discrepancies between local and global biodiversity trends. the study considers 161 long-term series from 115 marine, freshwater and terrestrial sites, mostly belonging to lter-europe, in 21 european countries, covering nine biogeographic regions, three realms and eight taxonomic groups. this wide scale, long-term study showed that a significant heterogeneity in the degree and direction of change in biodiversity metrics (i.e., the criteria used to ‘measure’ the level of biodiversity) exists at the continental spatial scale, demonstrating that biodiversity changes at local scale are often complex and cannot be easily generalized. going back to the three ecological strategies described by peters (2010), the papers by morabito et al. (2018), rogora et al. (2018) and pilotto et al. (2020) are examples of the “long” and the “broad”, which both represent essential components of the lter networks and of elterri, i.e., the wide scale systematic coverage of major terrestrial and aquatic environments, which can provide a whole-system multi-scale approach for comparative ecological analysis (mirtl et al., 2018). amongst the lter-italy aquatic sites, there are also many examples of the “deep”, meant as the development of different kinds of specific research questions and complementary studies, emerging from the long-term observations, and which provide valuable interpretative tools that enhance the informative significance of the long-term series themselves. i report here only two examples for all, concerning the lter site marechiara (lter-mc) in the gulf of naples (zingone et al., 2019) and the lake maggiore (tanentzap et al., 2020), two iconic lter sites as for the coastal marine environments and the large lakes, respectively (fig. 2). the paper by zingone et al. (2019) explicitly addresses how the lter local site, specifically, the ltermarechiara programme conducted with sampling 2 miles off downtown naples, proved to be relevant not only to trace plankton changes at different scales and under different environmental conditions across three decades, but this effort set a natural laboratory up, which has become a source of many inspiring voices for complementary research widening considerably the knowledge of the planktonic organisms and of the whole ecosystem functioning. a high number of phytoplankton species has been uncovered (both with microscopy-detailed taxonomic studies and biomolecular approaches), elucidating also the role of ‘crypticity’ in shaping natural diversity; the different phases in the life cycles of individual species have been finely described, with a consequent improvement of our ability to interpret seasonal and biogeographical patterns. at the other side of the ‘salinity border’, the long-time no nco mm er cia l u se on ly voices from the water 63 sseries of lake maggiore was used to thoroughly assess how the warming lake temperatures, up to 2 c, might have restructured the pelagic food webs from 1981 to 2008 (tanentzap et al., 2020). this paper addressed the thermal responses of the different trophic levels, associating temperatures with population dynamics, an issue whose evidence remain scarce in nature. as the lake warmed, the food web shifted, in terms of numerical abundance, towards predators occupying middle trophic positions, strengthening the top-down structuring of the food web. all the above-mentioned valuable studies addressed the plankton ecology and/or the functioning of ecosystem and their changes with the well-established cognitive approach, where we – as human beings in our role of scientists – observe the environment. field activities, critical data analysis and discussion, writing up the results: these, well-motivated and conscious, actions are all necessary and enjoyable activities taken by our rational mind, which allow obtaining valuable knowledge and objectively based information on how ecosystems work and change. however, i guess that we all have experienced another type of knowledge of the sites we are studying, based on different qualities that are connected with our emotions and with a peculiar kind of perception, which deeply bond us with nature, beyond scientific and rationale knowledge and words. this brings us to what is called “affective ecology” (barbiero, 2011 and 2017), a new discipline grounded on ecology and dealing, indeed, with our connecting with nature. affective ecology is epistemologically founded on two scientific hypotheses: the biophilia (wilson, 1984), which manifests the innate learning rules evolved in the human species to connect and govern our relationship with nature; and the theory of multiple intelligence (gardner, 1983), which individuate different manifestations of human intelligence, one of which can be defined as naturalist intelligence (gardner, 1999). in the scientific papers we write, there is usually (and understandably) virtually no place for an exploration of affective ecology and the qualities it brings, since these latter are considered to be just subjective impressions of our own, suitable for poetry perhaps, but not for a ‘business-as-usual’ approach to science. instead, cognitive ecology and affective ecology can act in synergy within one another, thus giving rise to a virtuous circle: on the one hand, establishing an affective connection with the natural world stimulates the desire to know nature at a fig. 2. lter – plankton sampling activities at the lter sites mc gulf of naples (left) and lake maggiore (right). pictures courtesy of maria grazia mazzocchi and gabriele tartari, respectively. no nco mm er cia l u se on ly a. pugnetti64 deeper level; conversely, a deeper ecological knowledge may stimulate a more intimate relationship between nature and the mankind (harding, 2011). are the lter-italy “voices” harmonized and open? based on the cross-site studies presented above, two major interconnected issues appear as shared at the european and global levels (mirtl et al., 2018): i) the heterogeneity of long-term series, for what concerns series lengths, survey schemes, and methodologies; and ii) the inadequate accessibility and availability of the dataset, which frequently lack appropriate data management plans and sharing procedures. these issues stem intrinsically from the most frequent establishment process of national networks at the european level, including the italian one, which were mainly driven by bottom-up efforts, i.e., making the best use of the existing long-term ecological activities and facilities, without any prior harmonization and agreement around data management procedures and policies. at the global and european levels both aspects are tackled with different tools and initiatives, namely by the adoption of: i) the ecological metadata language (eml, michener et al., 1997) as a common metadata language (vanderbilt et al., 2010); ii) the environmental thesaurus (envthes, http://vocabs.ceh.ac.uk/evn/tbl/envthes.evn; schentz et al., 2013), which provide a common and stable semantic backbone for documenting research products; and iii) the dynamic ecological information management system – site and dataset registry (deims-sdr), a research and monitoring site registry (https://www.deims.org/; wohner et al., 2019) that allows describing and identifying in situ sites and facilities (stations, sensors, datasets, persons), generating persistent, unique and resolvable identifiers for each site, and also documenting associated data linked to each site. besides, a core set of biotic and abiotic variables is under definition, for better comparisons across spatial scales and increase of the usage of lter data, based on the already existing frameworks and approaches at the global level (haase, 2018). these core set of recommended variables consider and link, currently, both the ecosystem integrity framework (mueller, 2005) and the essential biodiversity variables (ebv; pereira et al., 2013), within an integrated approach to be applied to terrestrial, freshwater, transitional and marine ecosystems. this approach could support holistic ecosystem research, ensuring that individual sites cover the most important features of ecosystems, promoting the comparability and interoperability of data between different sites and monitoring networks, and contributing to harmonize field research globally. to this respect, zilioli et al. (2019) tested the level of maturity of lter-italy at contributing to the ebvs’ framework by i) investigating how many and which categories of lter sites currently provide datasets suitable for measuring the two ebvs “species distribution” and “species abundance”; and ii) mining the information which is structured and publicly shared by site managers through deims-sdr. through this analysis, a first overview of ebvs monitored in lter-italy and the corresponding data management practices has been provided, as well as an assessment of the interoperability of the network with respect to other research organisations. the study evidenced a documented capacity to provide essential measures through the information system deimssdr, but at the same time the need to support the community and to optimise the ebvs’ information retrieval to improve the assessment and hence the effectiveness of lter as an observing system. the technical, cultural and social aspects of data sharing are considered as crucial in ilter, which thoroughly address them with a general and full agreement, in principle, on open data at the global scale (vanderbilt et al., 2015; vanderbilt and gaiser, 2017; shin et al., 2019). however, putting the open science principles into practice is still an issue in many of the national member networks and at the local level, calling for the development of practical case studies. one of them has been recently carried out in the lter-italy framework: it regarded the release, with open access, of a long term marine dataset of water quality and plankton data collected in the northern adriatic sea, in 50 years, from 1965 to 2015 (http://doi.org/10.5281/zenodo.3516717; acri et al., 2019, 2020). this goal was achieved by embracing an open vision extended to the whole research lifecycle (minelli et al., 2018a, 2018b): from research idea to results and data, from metadata to methods and software. open science actually involves a broad and comprehensive view, for making the various research steps accessible, reusable, and understandable. besides the most technical aspects, which can be found in the above-cited papers and that are not reported here, i wish to stress the cultural and human ones, which connected the lter voices beyond time and space. indeed, a heterogeneous working group, made by field ecologists and data management scientists as well as by retired researchers, worked together since the beginning and along the whole process, with the goal of sharing and harmonizing also the different experiences, needs, and points of view, and of demonstrating that a change of vision is possible, from “publishing as soon as possible” to “sharing and collaborating” (moedas, 2015). all the lteritaly scientists involved agreed with open science and they were willing to know better what open science actually could imply in practical terms. uncertainty to openly share the data was mainly due to the concern of getting proper acknowledgments to the data producers and about the extra time needed to accomplish the whole process, no nco mm er cia l u se on ly voices from the water 65 which was seen as a further workload, difficult to hold up. moreover, the opportunities of publishing datasets and data-papers were also poorly known and considered with scepticism, mainly for the low impacts on the professional careers. actually, indexing published datasets on web of science and the practice of data papers is quite a new one, which needs still to enter in the publication habits. working side by side in this multidisciplinary group allowed to express and examine honestly all the difficulties encountered, trying to find together the possible solutions and including, as essential parts and outcomes of the process itself, also the cultural resistances and practical challenges. the whole process required sharing, in a concrete and pragmatic way, visions, tools, and languages. one of the unexpected products was, for instance, a glossary (scovacricchi, 2020) initially started as a joke after each meeting, containing all the terms and acronyms that were used in the informatics field and were not understandable by field researchers. this glossary has been then organized, transformed into a report, openly accessible (http://doi.org/10.5281/zenodo.4302115), to be exploited as a tool to learn and share terminologies that often could hamper the integration between different expertise. finally, i wish to emphasize the contribution to lter studies provided by retired researchers, who revealed to be extremely precious, for the knowledge they owned about the oldest data, which they gathered personally, and for their unaltered memories, tightly connected with personal stories, which were reviving the data from the past. thanks to them, we became more aware of the human dimension behind the dataset and of the crucial importance of fostering open science also for allowing the creation of links among the past, the present and the future. this brings us back to the paper by magnuson (1990) and to his call for the responsibility in lter studies: transmitting openly the data and knowledge, inspired by the vision of continuity across generation, can actually move us from a pure cognitive experience to an enlivening and rewarding one. lter-italy voices in dialogue with society there is a wide and growing recognition that the environment must be viewed and studied as a social-ecological system, where ecological research becomes also a cultural process, not only a scientific one (haberl et al., 2006). the integration of social sciences in lter networks is a recognized priority (singh et al., 2013; mirtl et al., 2018, dick et al., 2018): socio-ecological research is conducted worldwide in lter networks, aiming at involving a broad stakeholder community in measurements in the co-design of investigation practices and also in the definition of research priorities (haberl et al., 2006; mauz et al., 2012; dick et al., 2018). the lter networks may therefore represent an appropriate and suitable context where new and different forms of communication and public participation and engagement could be experimented, sharing the voices of lter with the civil society in a mutual relationship. indeed, ecosystem and biodiversity too often remain abstractions and conceptualizations, unless new methodologies in science communication and education are developed, with new voices and words able to convey different perceptions, embedded in the language, in the culture, and in the daily experiences (monbiot, 2017; gray and colucci gray, 2018). aiming at making the public more familiar with the different lter ecosystems and with the lter vision and aims, lter-italy researchers planned and realized, from 2015 to 2019, the informal science-communication initiative called cammini lter (d’alelio et al., 2016; bergami et al., 2018; l’astorina et al., 2018a; pugnetti et al., 2019). during cammini, researches “walked, cycled and kayaked” along itineraries, which connected two or more lter sites, creating a physical and visible movement of researchers towards and with citizens, sharing informal events and communication activities, in close relationship and cooperation with the territories crossed, which were quite heterogeneous in size (from big towns to small villages) and audience (from school children to elderly people, from lay people to territorial managers, such as foresters, ecological and alpine guards, local environmental associations). thirteen trails were realized in 5 years, most of them involving lter aquatic sites (tab. 1; bergami et al., 2018) and letting spread the voice of plankton in very different context, across seas, lagoons, lakes, and also on the land. the initiatives realized in the trails covered most of the communication typologies mentioned in the literature (bergami et al., 2018; l’astorina et al., 2018b; pugnetti et al., 2019). the science communication system can be, indeed, compared someway to an ecosystem (davies and horst 2016), where various communication practices coexist, occupying diverse and self-consistent niches, all quite interconnected. in the cammini, traditional (e.g., press releases, public conferences, tweets and post on the social networks, reports on blogs) and experimental (fig. 3; e.g., sharing of the lter activities at the sites, with samplings of the different ecosystems’ components and microscopy sessions of plankton and benthos) activities were carried out along each trail, aiming at highlighting the relevance of lter in the territories crossed and the role of the institutions involved. to these activities, more participative and inclusive initiatives were added, such as those of citizen science (criscuolo et al., 2018a, 2018b), the biobliotz (petriccione, 2018) and the sea futuring tours (fig. 3; l’astorina et al., 2018c). these lasts consist in a new form of public engagement, aimed at fostering awareness and responsibility towards the surrounding environment, the sea in this case: secondary school students no nco mm er cia l u se on ly a. pugnetti66 of naples and taranto were invited and guided, through different cognitive and sensorial experiences, to imagine the future of the sea they lived by, the gulf of naples and the mar piccolo. the multifaceted communication activities and methods experienced during cammini reveals the heterogeneity of expertise, interests and expectations of lter scientists: i) experience new languages to make science most accessible and attractive; ii) activate synergies between expert and local knowledge in a mutual exchange; and iii) recover or strengthen the emotional and affective link with our work and with the natural environment. cammini can be considered as a part of a long-lasting tradition, where walking is considered the most intimate way to engage with the landscape, offering privileged insights and knowledge into both places and self (solnit, 2000), reinforcing the connection with the natural environment (maturana and varela, 1998; varela et al., 1991). in the physical gesture of “opening to the world” with the body and the senses (le breton, 2000), most researchers left behind, not only symbolically, the comfort zone they were used to (e.g. the desks, the laboratories, the mental schemes, the thematic congresses) to experiment informal and not mediated communication modalities, activating empathy with the people and places and overcoming the sense of separation and distance that often characterize the relations between science and society (l’astorina et al., 2018a). conclusive musings in this paper i evidenced some crucial aspects of lter-italy activities at the aquatic sites, with emphasis on plankton, which guiding voice brought me to deal different issues that have been and still are at stake in the current development of the national network and of fig. 3. examples of activities carried out during cammini lter. upper panel: scientist sharing with the public the sampling activities at the lter mountain site lake paione inferiore. lower panel: secondary school students engaged in the sea futuring tours at the lter site gulf of naples, gathering and analysing materials to imagine possible futures for the sea. pictures courtesy of antonio bergamino. no nco mm er cia l u se on ly voices from the water 67 elter-ri, at the european level: i) the relevance of wide scale comparative ecological analysis at different sites across ecological domains; ii) the possibility to develop specific research questions and complementary studies emerging from the long term observations at the lter sites; iii) the necessity to deal with methods and data heterogeneity, by identifying harmonizing frameworks and fostering open science; and iv) the possibility to find and pursue a change in our vision of the relations between science and society, starting with different ways of communicating, in an open and empathic way. most of these aspects are an expression of our scientific and cognitive approach to ecosystem study and they are all necessary to gaining knowledge and information on how ecosystems work and change through time. in this conclusive musings i wish to make a step further, emphasizing some voices that are much overlooked, and which whisper that the lter sites could be also the starting point where develop a deep connection with nature, entering in what is called the “soul of the place” (harding, 2011): places in which one may spend time getting to know them intimately with intuition and feeling, whilst using the rational mind to find out about their ecology and about how humans have interacted with these places over the years. the age we live in is characterized by uncertainties, complexity and by crucial environmental and societal challenges (millenium ecosystem assessment, 2005): according to many scholars, it is necessary, more than ever, to embrace a shift of perspectives for a transformative change (leiserowitz et al., 2005; ericson et al., 2014), which requires the elaboration of a deep and broadened form of subjectivity, in which humans identify also with non-humans, allowing to suspend the illusion of separateness (naess, 1984; 1995; carvalho, 2014; geiger et al., 2018; eisenstein, 2018). indeed, as evidenced by an increasing number of studies, scientific knowledge, new technology or governance alone cannot resolve the current ecological crisis and global sustainability challenges (wamsler and brink, 2018). they also require a broader cultural change and new approaches (esbjörn-hargens and zimmermann, 2009; dhiman and marques, 2016; wamsler and brink 2018) that integrate in every ecological study and action an inner dimension, which include emotions, identities and values (buchanan and kern, 2017; hedlund-de witt, 2011; ives et al., 2018, 2019). the inner world has been largely disregarded in scientific studies, but it is gaining increasing momentum in parallel scientific fields, such as neuroscience, ecopsychology and education. within this context, some contemplative practices, such as meditation, have been proved to be crucial to open pathways (mayer and frantz, 2004; panno, et al., 2017; geiger et al., 2018; wamsler, 2018), which could enable an expansion of empathy to include non-human subjects or, as inspiringly david abrams (1997) calls them, the “more than human world”. i decided not by chance to start this paper citing a famous sentence of the buddhist monk tich nath han, whose practices and insights aim at fostering the awareness of interconnectedness and “inter-being” (carvalho, 2014), since he expresses in a poetical and vibrant way the deep interconnections that are at the foundation of any ecological and lter study: interconnections among human beings, in a long chain that links the past to the present and to the future through lter studies, and among us humans and non-human beings that surround us, in intertwined and mutually dependent functions and processes (abrams, 1997). acknowledgments i think that every paper should start with a declaration of gratitude, deeply and openly appreciating the so many beings and events without those we simply could not be able to be here alive, willing to write and share our views and results. this is particularly true when speaking of lter, a field of activity that intrinsically connect us with a vast community of non-human and human beings along the past and the future. keeping this in mind and with my gratitude extending all around, i will here acknowledge explicitly just a few persons, among the many i am indebted to: caterina bergami, amelia de lazzari, alba l’astorina, and bruno petriccione, who accompanied me in the years with emotions, thoughts, reflections, inspiration, and friendship. this work was presented as a plenary invited lecture at the xxiv congress of the italian association of oceanology and limnology (bologna, 5-7 june 2019). i am grateful to domenico d’alelio for his accurate revision of the paper and for the time of his life he dedicated to it. finally, i wish to dedicate the paper to four dearest colleagues who passed away recently: lorenzo busetto, mario contesini, mario giordano and giuseppe morabito. corresponding author: alessandra.pugnetti@ismar.cnr.it key words: plankton, lter-italy, aquatic ecosystems, affective ecology, meditation received: 16 november 2020. accepted: 7 december 2020. this work is licensed under a creative commons attribution noncommercial 4.0 license (cc by-nc 4.0). ©copyright: the author(s), 2020 licensee pagepress, italy advances in oceanography and limnology, 2020; 11:9508 doi: 10.4081/aiol.2020.9508 no nco mm er cia l u se on ly a. pugnetti68 references abrams d, 1997. the spell of the sensuous. perception and language in a more-than-human world. vintage books. 352 pp. acri f, bastianini m, bernardi aubry f, boldrin a, camatti e, et al., 2019. lter northern adriatic sea (italy) marine data from 1965 to 2015 (version 3) [data set]. zenodo. doi: 10.5281/zenodo.3516717 acri f, bastianini m, bernardi aubry f, camatti e, boldrin a, et al., 2020. a long term (1965–2015) ecological marine database from the lter-italy site northern adriatic sea: plankton and oceanographic observations. earth syst. sci. data 12:215-230. barbalet j, 2002. science and emotions. sociol rev 50:s132-150. barbiero g, 2011. biophilia and gaia: two hypotheses for an affective ecology. j. biourbanism 26:1-27. barbiero g, 2014. affective ecology for sustainability. visions for sustainability, 1: 20-30. barbiero g, 2017. [ecologia affettiva].[book in italian]. mondadori, 228 pp. bergami c, l’astorina a, pugnetti a, 2018. [i cammini della rete lter-italia. il racconto dell’ecologia in cammino].[book in italian]. cnr edizioni. 180 pp. buchanan a, kern ml, 2017. the benefit mindset: the psychology of contribution and everyday leadership. int. j. wellbeing 7:1–11. cappello p, 2013. [questa libertà].[book in italian]. rizzoli editore. 180 pp. carvalho a, 2014. subjectivity, ecology and meditation – performing interconnectedness. subjectivity 7:131-150. criscuolo l, carrara p, oggioni a, pugnetti a, antoninetti m, 2018a. can vgi and mobile apps support long term ecological research? a test in remote areas of the alps. in: g. bordogna and p. carrara p. (eds.), mobile information systems leveraging volunteered geographic information for earth observation. earth systems data and models vol 4. springer, cham. criscuolo a, oggioni a, campanaro a, 2018b. [la citizen science nei cammini lter]. in: c. bergami, a. l’astorina and a. pugnetti a (eds.), [i cammini della rete lter-italia. il racconto dell’ecologia in cammino].[book in italian]. cnr edizioni. d’alelio d, 2016. the mesothalassia bike-tour: (re) discovering water by riding with scientists. limnol. oceanogr. bull. 25: 1-7. d’alelio d, 2020. [la microgiungla del mare].[book in italian]. hoepli editore. 177 pp. davies sr, horst m, 2016. science communication: culture, identity and citizenship. palgrave macmillan. dhiman s, marques j, 2016. spirituality and sustainability: new horizons and exemplary approaches. springer, cham. dick j, orenstein de, holzer jm, wohner c, achard al, andrews c, et al., 2018. what is socio-ecological research delivering? a literature survey across 25 international ltser platforms. sci. total environ. 622-623: 225-1240. eisenstein c, 2018. climate: a new story. north atlantic books. 306 pp. esbjörn-hargens s, zimmermann me, 2009. integral ecologyuniting multiple perspectives on the natural world. integral books, boston. geiger sm, otto s schrader u, 2018. mindfully green and healthy: an indirect path from mindfulness to ecological behavior. front. psychol. 8:2306. gardner h, 1983. frames of mind: the theory of multiple intelligences. new york, basic books. 440 pp. gardner h, 1999. intelligence reframed. new york, basic books. gray ds, colucci-gray l, 2018. laying down a path in walking: student teachers’ emerging ecological identities. environ. educ. res. 25:341-364. haase p, tonkin jd, stoll s, burkhard b, frenzel m, geijzendorffer ir, et al., 2018. the next generation of site-based long-term ecological monitoring: linking essential biodiversity variables and ecosystem integrity. sci. total environ. 613-614: 376-1384. haberl h, winiwarter v, andersson k, ayres ru, boone c, castillo a, et al., 2006. from lter to ltser: conceptualizing the socioeconomic dimension of long-term socioecological research. ecol. soc. 11:13. harding s, 2011. animate earth. green books: 258 pp. hedlund-de witt a, 2011. the rising culture and worldview of contemporary spirituality: a sociological study of potentials and pitfalls for sustainable development. ecol. econ. 70:1057-1065. hensen v, 1887). [ueber die bestimmung des plankton’s oder des im meere treibenden materials an pflanzen und thieren].[book in german]. kiel: schmidt & klaunig. ives cd, abson dj, von wehrden h, dorninger c, klaniecki k, fischer j, 2018. reconnecting with nature for sustainability. sustain. sci.13:1389-1397. ives cd, freeth r, fischer j, 2019. inside-out sustainability: the neglect of inner worlds. ambio 49:208-217. koppman s, cain cl, leahey e, 2015. the joy of science: disciplinary diversity in emotional accounts. sci. technol. hum. values 40:30-70 l’astorina a, bergami c, d’alelio d, dattolo e, pugnetti a, 2018a. what is at stake for scientists when communicating ecology? insight from the informal communication initiative “cammini lter”. visions for sustainability 10:19-37. l’astorina a, bergami c, pugnetti a, 2018b. [raccontare l’ecologia in cammino per aprirsi al mondo]. in: c. bergami, a. l’astorina and a. pugnetti a (eds.), [i cammini della rete lter-italia. il racconto dell’ecologia in cammino].[book in italian]. cnr edizioni. l’astorina a, pelusi a, petrocelli a, portacci g, rubino f, 2018c. i sea futuring tours nei cammini lter. in: c. bergami, a. l’astorina and a. pugnetti a (eds.), [i cammini della rete lter-italia. il racconto dell’ecologia in cammino].[book in italian]. cnr edizioni. le breton d, 2000. [eloge de la marche].[book in french]. paris, editions métailé. magnuson jj, 1990. long-term ecological research and the invisible present. bioscience 40:495-508. maturana h, davila x, 2009. preface to a beautiful book. in: j. cull (ed.), the circularity of life – an essential shift for sustainability. j. cull: 142 pp. maturana hr, varela fj, 1998. the tree of knowledge: the biological roots of human understanding. revised edition. boston: shambhala publications. 269 pp. mauz i, peltola t, granjou c, van bommel s, buijs a, 2012. no nco mm er cia l u se on ly voices from the water 69 how scientific visions matter: insights from three long-term socio-ecological research (ltser) platforms under construction in europe. environm. sci. policy 19:90-99. mayer fs, frantz cm, 2004. the connectedness to nature scale: a measure of individuals’ feeling in community with nature. j. environ. psychol. 24:503–515. michener wk, brunt jw, helly jj, kirchner tb, stafford sg, 1997. nongeospatial metadata for the ecological sciences. ecol. appl. 7:330–342. millennium ecosystem assessment, 2005. ecosystems and human well-being, volume 1. current state and trends. washington: island press. minelli a, bergami c, oggioni a, pugnetti a, 2018a. [il mare, la ricerca ecologica a lungo termine e la scienza aperta: lavori in corso]. in: l’astorina a, di fiore m, 2018. [scienziati in affanno? ricerca e innovazione responsabili (rri) in teoria e nelle pratiche].[book in italian]. cnr edizioni. minelli a, oggioni a, pugnetti a, sarretta a, bastianini m, bergami c, et al., 2018b. the project econaos: vision and practice towards an open approach in the northern adriatic sea ecological observatory. res. ideas outcomes 4:e24224. mirtl m, borer e, djukic i, forsius m, haubold h, hugo w, et al., 2018. genesis, goals and achievements of long-term ecological research at the global scale: a critical review of ilter and future directions. sci. total environ. 626:14391469. moedas c, 2015. freedom is absolutely necessary for scientific progress. speech at european commission. brussels, 26 january 2015. accessed: 24 august 2020. retrieved from: https://ec.europa.eu/commission/commissioners/20142019/moedas/announcements/freedom absolutely-necessary-scientific-progress_en mollenhauer h, kasner m, haase p, peterseil j, wohner c, frenzel m, et al., 2018. long-term environmental monitoring infrastructures in europe: observations, measurements, scales, and socio-ecological representativeness. sci. total environ. 624:968–978. monbiot g, 2017. forget ‘the environment’: we need new words to convey life’s wonders. the guardian, august 9. accessed: 23 october 2020. retrieved from: https://www.theguardian.com/commentisfree/2017/aug/09/forget-the-environment-new-words-lifes-wonders-language morabito g, mazzocchi mg, salmaso n, zingone a, bergami c, flaim g, et al., 2018. plankton dynamics across the freshwater, transitional and marine research sites of the lteritaly network. patterns, fluctuations, drivers. sci. total environ. 627:373-387. muller f, 2005. indicating ecosystem and landscape organisation. ecol. indic. 5:280–294. naess a, 1984. identification as a source of deep ecological attitudes, p. 256–270. in: m. tobias (ed.), deep ecology. san diego: avant books. naess a, 1995. self-realization: an ecological approach to being in the world, p. 13-30. in: a. drengon and y. inoue (eds.), the deep ecology movement. berkeley, north atlantic books. panno a, giacomantonio m, carrus g, maricchiolo f, pirchio s, mannetti l, 2017. mindfulness, pro-environmental behavior, and belief in climate change: the mediating role of social dominance. environ. behav. 50:864-888. pereira hm, ferrier s, walters m, geller gn, jongman rhg, scholes rj, et al., 2013. essential biodiversity variables. science 339:277-278. peters dbc, 2010. accessible ecology: synthesis of the long, deep, and broad. trends ecol. evol. 25:592–601. petriccione b, 2018. i bioblitz nei cammini lter. in: c. bergami, a. l’astorina and a. pugnetti a (eds.), [i cammini della rete lter-italia. il racconto dell’ecologia in cammino].[book in italian]. pilotto f, kühn i, adrian r, alber r, alignier a, andrews c, et al., 2020. meta-analysis of multidecadal biodiversity trends in europe. nature commun. 11:3486. pugnetti a, de lazzari a, bergami c, l’astorina, 2019. [il dialogo dei ricercatori con la società civile. esperienze e riflessioni dai cammini della rete di ricerca ecologica a lungo termine lter-italia].[article in italian]. museologia scientifica memorie 20:127-134. rogora m, frate l, carranza ml, freppaz m, stanisci a, bertani i, et al., 2018. assessment of climate change effects on mountain ecosystems through a cross-site analysis in the alps and apennines. sci. total environ. 624:1429-1442. schentz h. peterseil j, bertrand n, 2013. envthes-interlinked thesaurus for long term ecological research, monitoring, and experiments, p. 824-832. proceedings enviroinfo 2013: environmental informatics and renewable energies. aachen, shaker verlag. scovacricchi t, 2020. a glossary for open science, open geospatial and data curation. zenodo. http://doi.org/10.5281/zenodo.4302115 shin n, shibata h, osawa t, yamakita t, nakamura m, kenta k, 2019. toward more data publication of long-term ecological observations. ecol. res 35:700-707. singh sj, haberl h, chertow mr, mirtl m, schmid m, 2013. long term socio-ecological research: studies in society-nature interactions across spatial and temporal scales. cham, springer. pp. 588. solnit r, 2000. wanderlust: a history of walking. new york, penguin books. tanentzap aj, morabito g, volta p, rogora m, yan nd, manca m, 2020. climate warming restructures an aquatic food web over 28 years. global change biol. doi: 10.1111/gcb.15347 vanderbilt k, gaiser e, 2017. the international long term ecological research network: a platform for collaboration. ecosphere 8:e01697. vanderbilt kl, blankman d, guo x, he h, lin c-c, lu s-s, et al., 2010. a multilingual metadata catalog for the ilter: issues and approaches. ecol. inform. 5:187-193. vanderbilt kl, lin c-c, lu s-s, kassim ar, he h, guo x, et al., 2015. fostering ecological data sharing: collaborations in the international long term ecological research network. ecosphere 6:1-18. varela fj, thompson e, rosch e, 1991. the embodied mind: cognitive science and human experience. cambridge, mit press. wamsler c, 2018. mind the gap: the role of mindfulness in adapting to increasing risk and climate change. sustain. sci. 13:1121-1135. wamsler w, brink e, 2018. mindsets for sustainability: exploring the link between mindfulness and sustainable climate adaptation. ecol. econom. 151:55-61. no nco mm er cia l u se on ly a. pugnetti70 wilson eo, 1984. biophilia. cambridge, harvard university press. wohner c, peterseil j, poursanidis d, kliment t, wilson m, mirtl m, chrysoulakis n, 2019. deims-sdr – a web portal to document research sites and their associated data. ecol. inform. 51:15-24. zilioli m, oggioni a, tagliolato p, pugnetti a, carrara p, 2019. feeding essential biodiversity variables (ebvs): actual and potential contributions from lter-italy. nature conservation 34:477-503. zingone a, d’alelio d, mazzocchi mg, montresor m, sarno d, lter-mc team. 2019. time series and beyond: multifaceted plankton research at a marine mediterranean lter site. nature conservation 34:273-310. no nco mm er cia l u se on ly layout 1 introduction the functioning of riverine ecosystems is the result of complex interactions among multiple abiotic and biotic components which, altogether, contribute to the establishment of specific and dynamic mosaics of aquatic and terrestrial habitats. the variability of flows in terms of magnitude, frequency, duration, timing and rate of change, controls river ecosystem functioning, by preserving their integrity and allowing the persistence of native species and natural habitats, as well as the development of the human society (poff et al., 1997; bunn and arthington, 2002; acreman, 2016). nowadays, the increasing intensification of human activities is the primary factor of rivers natural flow regimes alteration (peñas et al., 2016; horne et al., 2017). as a consequence of a growing human population and associated economic development, many anthropogenic interventions have led to significant degradation of more than 60% of the world’s rivers (nilsson et al., 2005). global water withdrawals increase by about 10% every 10 years, with this rate being much higher in developing regions (organisation for economic co-operation and development, 2012). the exploitation of water resources for urban, agricultural and industrial purposes, together with changes in land use, transport and energy production negatively affects both the hydrology (reduced residual water) and the morphology of rivers (longitudinal and lateral connectivity interruption) (lake, 2007; döll et al., 2009). in parallel, climate change is also influencing rivers’ flow regime (barnett et al., 2005; blöschl et al., 2017; pletterbauer et al., 2018). indeed, the outflow of rivers is directly influenced by long-term changes in both precipitation and temperature trends (dudgeon et al., 2006). variations in the amount of precipitations associated with global change can modify the hydrological cycles, which in turn, influence habitats availability for many aquatic organisms (filipe et al., 2013). temperature variations affect the timing of the outflow and modify the chemical-physical features of the aquatic environment, ultimately altering ecological and biological processes (pletterbauer et al., 2018). furthermore, climate change can cause a deep modification of human usages of rivers’ water or directly interact with human pressures. for instance, in many regions of the world, due to the increasing water scarcity, a growing number of water management measures, such as impoundments, diversion weirs and dams, are being adopted, contributing significantly to lowering rivers’ habitats quality (postel, 1998; magilligan et al., 2005; zhang et al., 2015). a well-established approach aimed to preserve freshwater ecosystems and river ecological status relies on the definition of a sustainable environmental flow review implementation of the eu ecological flow policy in italy with a focus on sardinia davide moccia,1* luca salvadori,2 simone ferrari,2 alessandra carucci,2 antonio pusceddu1 1department of life and environmental sciences, university of cagliari, via tommaso fiorelli 1, 09126 cagliari; 2department of civil-environmental engineering and architecture, university of cagliari, via tommaso fiorelli 1, 09126 cagliari, italy abstract river ecosystems are characterised by a naturally high level of hydrodynamic perturbations which create aquatic-terrestrial habitats indispensable for many species, as well as for the human beings’ welfare. environmental degradation and habitat loss caused by increasing anthropogenic pressures and global change affect freshwater aquatic ecosystems worldwide and have caused changes in water flow regimes and channels morphologies. these, in turn, decreased the natural flow capacity and reduced habitat availability, thus causing severe degradation of rivers’ ecological integrity. the ecological flow (e-flow) is commonly intended as the quantity, timing, duration, frequency and quality of water flows required to sustain freshwater, estuarine and near shore ecosystems and the human livelihoods and well being. maintaining the e-flow represents a potential tool for restoring and managing river ecosystems, to preserve the autochthonous living communities, along with environmental services and cultural/societal values. in the last decade, methods for the determination of the e-flow in european rivers moved from a simply hydrological approach towards establishing a linkage between the hydrological regime and the good ecological status (ges) of the water bodies, as identified by the european water framework directive (wfd; 2000/60/ec). each member state is required to implement and integrate into the river basin management plans (rbmp) a methodology for the determination of the e-flow, ensuring that rivers can achieve and maintain the ges. the competent river basin authorities have thus to ascertain whether national methodologies can be applied to different river typologies and basin environment characteristics. in this context, we narratively review the e-flow assessments in the heterogeneous italian territory, in particular on a water scant region such as sardinia, by analysing laws, guidelines and focusing on study cases conducted with micro and meso-scale hydraulic-habitat approaches. in the sight of a more ecological-based application of national e-flow policy, we suggest that meso-habitat methods provide a valuable tool to overcome several limitations of current e-flow implementation in the italian territory. however, to face future challenges, such as climate change adaptation, we stress the need for further experimental studies to update water management plans with greater attention for nature conservation. no nco mm er cia l u se on ly implementation of the eu ecological flow policy in italy with a focus on sardinia 23 release from dams and other types of derivation (acreman et al., 2016). the environmental flow (e-flow) is commonly intended as the quantity, timing, duration, frequency and quality of water flows required to sustain freshwater, estuarine and near shore ecosystems and the human livelihoods and well being (tharme, 2000). the concept of e-flow has been discussed for more than 40 years (tharme, 2003; acreman and dunbar, 2004; king and brown, 2006; poff and matthews, 2013; acreman, 2016) and it is now worldwide recognized by several national and regional water protection policies, e.g. the european water framework directive (european commission, 2015), the south african national water act (forslund et al., 2009) and the brisbane declaration (arthington et al., 2018). many countries have incorporated e-flow provisions as they have updated water policy and laws (le quesne et al., 2010). over the last few years, there has been a growing consensus about the need of science-based approaches and requirements for the assessment of e-flows in order to achieve successful management of freshwater ecosystems, conciliating natural habitats conservation with the supply of freshwater for human usage (acreman and dunbar, 2004). the achievement of such a trade-off can become more critical in the sights of modified hydrological conditions that will result as a consequence of current climate change (tonkin et al., 2019). tharme (2003) identified >200 methods to assess the water requirements of aquatic species and habitats, and support e-flows management practices to achieve both ecological and social targets (see also acreman and dunbar 2004). however, to date, there are still several limiting factors for the implementation of successful eflow policy, due, mainly, to limitations in institutional capacity, scientific knowledge and monitoring resources. these have been summarized by le quesne et al., (2010) in four crucial issues: i) the scarcity of knowledge about sites and speciesspecific e-flows requirements; ii) the lack of political decision and effort to recognizing a need to maintain sufficient water in rivers and wetland systems; iii) imperfect understanding of the practical issues and costs of implementation to achieve the e-flow regime; iv) the lack of a design and maintenance of robust monitoring systems to assess the ecological and other outcomes of environmental watering in an adaptive management framework. in order to overcome these general issues, research and monitoring of e-flows’ adjustments is often required (richter, 2009; zang et al., 2012), as well as, the synthesizing of knowledge and experience gained from individual case studies (arthington et al., 2004; poff et al., 2010). in fact, regional and local approaches have a fundamental role in the understanding e-flow methods limitations. they will be crucial for the implementation of an applicable methodology to define e-flows by river basin authorities. this is particularly needed in a context where heterogeneous climatic regions (e.g. alpine, mediterranean) occur and a homogeneous and more ecological-based methodology for the definition of the eflow is requested. in this context, we review the status of e-flow implementation in italy, with a focus on a region with high water scarcity such as sardinia, by analyzing laws, guidelines and study cases conducted with hydraulic-habitat approaches, both at micro and mesoscale. the final aim of this review is to highlight current challenges and limitation that make difficult the application and adaptation of a common e-flows methodology and give cues for the selection of a suitable habitat simulation models that could be integrated to overcome these limits. the european guidance on e-flows nowadays, in europe, the most significant pressures causing failure to achieve the good ecological status (ges) are the hydro-morphological ones (affecting ca. 40 % of surface water bodies, with the highest proportion reported for rivers and transitional waters) (e.e.a. report n. 7/2018). the largest proportion (26%) of water bodies is affected by physical alterations in the channel, bed, riparian zone, whereas up to 24% of the investigated water bodies are affected by the presence of structures, such as dams/barriers and locks, that alter the rivers’ longitudinal continuity (e.e.a. report n. 7/2018). since 2012, to implement e-flows management in the river basin management plans the european commission (ec) has asked the member states (mss) to focus on pressures affecting the hydrological status of water bodies (schmidt and benítez sanz, 2012). however, the lack of harmonized methodologies and, consequently, of sufficient or sufficiently consistent data for the definition of e-flows’ assessment by each ms has raised several operational and interpretation difficulties (acreman and ferguson, 2010). for this reason, in 2015, the ec released the e-flow guidance document (ecological flows in the implementation of the water framework directive; wfd cis guidance document no. 31). in this document, the definition of e-flow is provided within the context of the water framework directive (wfd) as “a hydrological regime consistent with the achievement of the environmental objectives of the wfd in natural surface water bodies as mentioned in article 4(1)” and is interpreted as the “amount of water required for the aquatic ecosystem to continue to thrive and provide the services we rely upon”. once terminology has been clarified, concern has then arisen on how to no nco mm er cia l u se on ly d. moccia et al.24 calculate e-flows in different environmental contexts. the document did not report a full protocol for the implementation of e-flows nor led to a uniform e-flow implementation strategy, instead encouraged mss to share the knowledge obtained so far to evaluate site-specific eflows implementation conditions (specific environmental values or ecosystem services) and to promote the e-flow definition in the wfd planning process, not considering it as a separate problem. in the guidance document, the e-flow quantification methods are provided and classified in three major categories (hydrological, hydraulic-habitat and holistic methodologies, table 1), also according to costeffectiveness, time-efficiency, complexity and focus (dyson et al., 2003; tharme, 2003; arthington et al., 2004; richter et al., 2006; king et al., 2008). hydrological methods are based on the analysis of the natural flow regimes derived from historical streamflow data (mean annual flows, monthly flows, high/low flows or more complex hydrological indices) that can be existing or simulated. these methods provide an overall baseline flow that aims to conserve the natural flow integrity of a stream and maintain native habitat and species. hydrological methods are still the most widely used approaches (european commission, 2015) and, often, are considered as initial analyses in the preliminary phases for supporting more complex methods. however, pure hydrology-based methodologies for e-flows assessment incompletely capture habitat dynamics and ecological responses of aquatic and riparian fauna. these methods fail to account for flow interactions with the channel morphology and its variation in space or modifications over time. furthermore, in the absence or lack of streamflow gauging stations, hydrological modelling of streamflow data series requires quite a lot of expertise, can be time-consuming and may produce significant uncertainties when predicting magnitude, frequency and timing of extreme events (low and high flows). hydraulic/habitat simulation methods are more sophisticated approaches in which habitat availability for aquatic species and their vital stages are expressed as a combination of hydro-morphological features and their ecological preferences. habitat modelling software such as telemac (galland, 1991), phabsim (usgs, 2001), rhyhabsim (jowett, 2010), casimir (schneider et al., 2010) and river 2d (steffler and blackburn, 2002) allow this type of accurate site and species characterisation at the micro-scale, while mesohabsim (parasiewicz, 2001, 2007), mesocasimir (schneider et al., 2001) and rhm (maddock et al., 2001) at the meso-scale. hydraulichabitat methods are often considered more accurate than hydrological ones since they complement them by incorporating flow-dependent and morphology-dependent ecological data, such as the occurrence of wetted areas and the connectivity between them, local hydrodynamic conditions of depth and flow velocity, sediment distribution and composition, and the presence of shelters and refuges for the fauna. however, they may require a considerable amount of fieldwork and expertise to collect the hydro morphological and biological data for model calibration. holistic methods are based on a global vision of the riverine ecosystem and encompass all major components or attributes of the river system. these methods require multidisciplinary inputs from different river scientists that integrate data, model predictions and expert knowledge (tharme, 1996; arthington, 1998; tharme, 2000; king et al., 2008). they consider both spatial and temporal table 1. list of e-flow approaches and methodologies in different countries and contexts. approach method source hydrological tennant (montana) tennant (1976) aquatic base flow caissie and el-jabi (1995) median monthly flow caissie and el-jabi (1995) range of variability approach ritcher et al. (1997) sustainability boundary approach and presumptive ritcher et al. (1997; 2011) hydraulic/habitat simulation phabsim (physical habitat simulation) bovee (1982) (modelling) methods mesohabsim parasiewicz (2001) evha (evaluation de habitat) ginot (1995) rss (river system simulator) alfredsen et al. (1995) casimir (computer aided simulation of habitat in regulated streams) jorde (2006) river 2d blackburn and steffler (2003) mesocasimir eisner et al. (2005) holistic frameworks generalized habitat models (e.g. stathab) lamouroux and jowett building block method (bbm) tharme and king (1998) drift (downstream response to imposed flow transformation) king et al. (2003) benchmarking arthington (1998; et al., 2006) eloha poff et al. (2010) no nco mm er cia l u se on ly implementation of the eu ecological flow policy in italy with a focus on sardinia 25 variations, allowing stakeholders and decision-makers to establish an acceptable risk as a balance among ecological goals, economic costs and scientific uncertainties. two different main categories of holistic methods exist: bottomup approaches, i.e. the building block methodology (king and louw, 1998), and top-down approaches, i.e. drift (downstream response to imposed flow transformation, king et al., 2003) and eloha (ecological limits of hydrological alteration; poff et al., 2010), that define eflows in terms of acceptable degrees of departure from the natural (or another reference state) flow regime and, therefore, require accurate quantification of ecological response to stress. to integrate the methodologies described above and to define a stepwise approach, the european e-flows framework, following the global trends, indicates three study levels: i) a preliminary risk analysis to define water bodies at risk of not achieving the wfd objectives because of hydrological alteration; ii) the establishment of an extensive hydrological monitoring network able to provide a good data collection in different water release situations to improve the usage of hydrological/habitat approaches in advanced steps of the analysis; iii) once the results of the experimental analysis are elaborated, the determination of a water-body typespecific environmental flow assessment (efa), together with an assessment of the gap between current and reference state (gap analysis), to guide specific measures. e-flow assessment in italy the italian territory overlooks the mediterranean sea and is characterised by a complex geo-morphological structure, with 51% of the country dominated by high mountains chains (the alps and the apennines), 29% covered by hills, and 20% occupied by plains (surian et al., 2003). such heterogeneous geo-morphological system interacts with the variability of the atmospheric circulation patterns making this mediterranean region one of the most vulnerable to future climate change (surian et al., 2003; giorgi, 2006). recent studies have predicted an increase in mean temperature of around 1.52.0°c in all seasons until 2050 (tomozeiu et al., 2017), together with an expected slight decrease of the mean annual rainfalls, and an intensification of extreme rainfall events (brunetti et al., 2001; giorgi and lionello, 2008; coppola and giorgi, 2010; lionello et al., 2012). additionally, scientific literature tends to agree that in spring and summer, change in the climate patterns is predicted to cause an increase in the frequency and intensity of droughts (spinoni et al., 2017). in this context, water demand is continuously increasing to satisfy the growing socio-economic needs and justified the construction of a significant number of large reservoirs in the italian territory, placing italy in the fifth place among the european member states with the highest number of dams (542, icold, 2007). additionally, in the last century, most of the italian rivers have experienced a considerable morphological change in terms of bed-level lowering (commonly of the order of 3–4 m), channel narrowing (in some cases up to 50%) and changes in channel pattern, due to sediment extraction, weirs and channelization (surian et al., 2003). such increasing impacts caused by these hydromorphologic pressures, coupled with the current climate change, have led to significant alteration of the hydrology of italian rivers, with consequent spatial and temporal reduction of fluvial habitats, accompanied by a decline of the environmental status and biodiversity loss (dudgeon et al., 2006; feyen et al., 2009; carlisle et al. 2010; filipe et al., 2013; blöschl et al., 2017; de girolamo et al., 2017; horne et al., 2017; spinoni et al., 2017). in 1989, with the enactment of the italian law n.183, the hydrological regime relevance was recognized by the italian government, which established the first legal act, including the qualitative definition of the e-flow. the concept of e-flow was initially associated to the idea of a “minimum constant flow” to be guaranteed in riverbeds, to promote a balance between human needs and the natural requirements of the riverine ecosystem (law 183/1989). the ministerial guidelines for the definition of the e-flow were issued only in 2004, and the subsequent legislative decree n.152/2006 introduced the concept of “minimum vital flow” (hereafter mvf), as the instant outflow needed to be preserved in downstream water diversions in order to conserve the physical (morphological, hydrological and hydraulic), physicalchemical (water quality) and biological features of natural riverine ecosystems. to date, the general formulas adopted among the italian districts by the river basin authorities, are based on simple hydrological formulas, mostly considering the percentage of mean annual or monthly flow, corrected through several coefficients taking into account different environmental aspects (table 2). these formulas are the most exploited in italy for their ease of application and cheapness. however, although this was the first step towards the implementation of a national rule to determinate the e-flow, three main problematic elements emerged: i) the absence of a univocal rule regarding the determination of the key hydrological parameters (e.g., annual average flow, the average flow of a specified period, the average flow of specific months); ii) the presence of significant diversity among calculation no nco mm er cia l u se on ly d. moccia et al.26 formulas resulting in territorial values of mvf notably heterogeneous, even within the same river basin district (when the rivers pass through different regions); iii) the absence of an evident and robust correlation between the mvf values and the environmental status of the water bodies classified by the indices required by legislative table 2. list of formulas adopted in different italian regions for assessing the minimal vital flow (mvf) in basin managements plans (2015-2021). regions formula legend po mvf=k × qmeda × s × m × z × a × t k: experimental parameter determined for individual hydrographical areas piemonte qmeda: specific annual average flow per unit of area of the basin (in l/s km2) liguria s: surface of the natural catchment area (kmq) toscana m: morphologic parameter valle d’aosta z: greater (n, f, q) emilia romagna n: coefficient of naturalness f: coefficient of fruition q: parameter concerning the quality of river water a: parameter concerning the interaction between surface waters and groundwater r: monthly hydrological modulation (class) t: time flow modulation. marche mvf(monthly)=[(k0-2,24 × 10-5 × s)] × qm) × k0: 0,075 (m × z × a × t) s: surface of the natural catchment area qm: average annual natural flow regime in the considered section m: morphologic parameter z: greater (n, f, q) n: current naturalness f: fruition parameter q: parameter concerning the quality of river water a: parameter concerning the interaction between surface waters and groundwater t: time flow modulation. provincia autonoma mvf(monthly) = mvf × sup × s × p × a × g × pe × r sup: absolute surface of the natural catchment area di trento s: surface of the natural catchment area p: monthly precipitations a: average altitude of the catchment area pe: basin and riverbed permeability g: riverbed geomorphology r: monthly hydrological modulation (class) veneto mvf= [(kbiol) +knat) × 207 × s0,8 × qmedia × 10-6 kbiol index of biological criticality knat: criticality naturalness index s: surface of the natural catchment area (km2) qmedia: average annual flow (l/s km2) friuli venezia giulia qmvf= k × t × p × m × qmedia qmedia: average annual flow to the section concerned k: protection level t: timing coefficient p: coefficient of naturalness and tourism m: coefficient of seasonal modulation abruzzo mvf = q × kbiol = (k1 × qrusc +k2 × qacqm + k3 × qrusc: runoff flow qsorg) × (kiif+kmorf+ki.b.e.+kmorf) qacqm: flow due to the emergence of minor aquifers qsorg: flow due to the emergence of significant groundwater bodies k1, k2, k3 are three coefficients that allow to weigh independently the contribution of each of the flow components defined above kiff: coefficient fluvial functional kmorf: morphologic coefficient kitt: ichthyofauna coefficient sicily mvf=10% (qmed) qmed: average annual flow sardinia mvf= 10% (qn) [in use] qn: natural flow mvf = qn (t,f) × k [under experimentation] qn (t,f); natural flow as a function of monthly mean rainfall k: coefficient related to hydro-morphological parameters no nco mm er cia l u se on ly implementation of the eu ecological flow policy in italy with a focus on sardinia 27 decree 152/06. to overcome the above-highlighted issues, in 2017 the italian ministry of the environment and protection of the territory and the sea (mattm) approved the directorial decree n. 30/sta (dd 30/sta hereafter), to update and provide homogeneous and scientifically advanced methodologies for the definition of the ecological outflow throughout the national territory, as requested by the european commission (cis guidance document n. 31/2015). implementation of hydraulic-habitat simulation models in italy within the italian territory, a variety of hydraulichabitat approaches have been tested to enhance the hydrological based e-flow assessments. the meso-scale or meso-habitat models were used in different italian regions for several species of fish. the geomorphic unit (or mesohabitat) generally ranges between about 1 m and 100 m (hauer et al., 2009; parasiewicz et al., 2013). this resolution allows the use of a wide range of environmental variables for the description of the fluvial habitat, obtaining a biological analysis of the life cycle of a single species or several communities, such a fish (gosselin et al., 2012; wilkes et al., 2015). the mesohabitat scale approach includes habitat suitability criteria to identify habitat characteristics in relation with different water regimes needed for the presence of particular species, for a particular biologically relevant periods of the year (e.g., spawning), or for particular endangered or threatened species (vezza et al., 2012). the mesohabsim model was used to describe habitat distribution and e-flow requirements needed to support local populations of the endangered white-clawed crayfish (austropotamobius pallipes) in lombardy region and in the gran sasso and laga national park mountainous streams (abruzzo region). with the use of a meso-habitat scale approach and time series analysis, detailed schemes of flow management were defined to represent habitat changes over time and to identify stress conditions for a. pallipes, created by the persistent limitations in habitat availability (vezza et al., 2016). the habitat suitability model mesohabsim, based on the statistic models random forest and logistic regression, was also used to estimate habitat preferences for the bullhead fish (cottus gobio) in alpine streams (vezza et al., 2014a). the distribution of bullhead fish in various habitat was also predicted with a mesohabsim model created with a deductive model, based on conditional habitat suitability criteria derived from expert knowledge and literature information (adamczyk et al., 2019). furthermore, the meso-habitat suitability model was used to quantify habitat alteration and to define e-flow schemes in high gradient streams using biological and hydro-morphological data from 40 study sites located within the mountainous areas of different regions, such as aosta valley, piemonte and liguria (northwest italy) (vezza et al., 2014b). moreover, the habitat meso-scale approach has been used to identify the optimal regime flow considering the limiting factors’ approach and the assessment of basic ecological relationships. in a study conducted in the serio river (northern italy), densityenvironmental variables relationships for three different life stages of the brown trout were investigated to select a range of flows able to preserve most of the physical habitat for all of the trout life stages. the results highlighted that the main factors limiting the trout densities were water velocity, substrate characteristics and availability of refugia from predation. this study also estimated the effect of varying discharge flows on macroinvertebrate biomass, and it was then used to identify an optimal flow maximizing either habitat and prey availability (fornaroli et al., 2013). in the aosta valley (north-western italy), the indices of river habitat integrity derived from the mesohabsim methodology were also used to demonstrate that the application of meso-scale habitat models allows the assessment of hydro-morphological spatio-temporal alterations of habitat’s structure and that these models and indices can be adequately integrated into a multi-criteria analysis framework, supporting decision-making processes by the competent authorities (vassoney et al., 2019). another type of hydraulic-habitat methodologies for estimating the e-flow assessment relies on the microhabitat approach. such an approach is used when the scale of investigation zooms in at site-specific or species-specific level, for example, when the focus switch on the conservation of one or few targeted species. this approach requires the acquisition of precise knowledge and data on the relationships between specific flow conditions and the life cycle requirements of the targeted species. such data are essential to create habitat suitability curves (hsc) and apply microhabitat-hydrological approaches to estimate a suitable e-flows tailored at the site-specific scale. however, in the literature, habitat suitability curves are available only for a few species, and the availability of experimental curves on the italian national territory is still very limited (vismara et al., 2001). another important limitation of the use of micro-habitat approaches is their dependence on hydraulic simulation models (1d or 2d), which are difficult to be used, or even inappropriate, in high gradient streams (vezza et al., 2014), in intermittent streams (acuña et al., 2020), in lowland macrophyte-dominated rivers (hearne et al., 1994) and in rivers with winter ice (alfredsen and tesaker 2002). experimental studies, using the microhabitat approach, have been mostly carried out to correct no nco mm er cia l u se on ly d. moccia et al.28 the suitability of flow regimes released by the basin river authorities’ formulas for several target species. the micro-habitat methodology phabsim was used by the marecchia-conca basin river interregional authority (salmoiraghi and morolli, 2007) to determine the value of the optimal flow needed to guarantee the functionality of the river system and the life of its organisms in several sections of the marecchia river and its tributary senatello (as a contribution to the partial basin plan for the hydrological instability). this study provided evidence of the need to correct the e-flows previously established to ensure the optimal habitat availability for the trout (salmo trutta), the barbel (barbus plebejus), and the chub (leuciscus cephalus) with those obtained from the application of the phabsim model. the micro-habitat approach, instream flow incremental methodology (ifim, bovee et al., 1998) was applied in the taro river (northern italy) to estimate the e-flow for three fish species (barbus plebejus, chondrostoma genei and leuciscus souffia) and their vital stages (young, adults and reproductive stage). once again, the results highlighted how the value of the e-flow, adopted so far for the taro river, calculated according to the general formula proposed by the basin authority for the po river, did not guarantee the preservation of any of the species analyzed in the study case. the same study also determined two different flow thresholds for each species: a “critical” threshold that indicates the flow in which species can “survive”, and an “ecological flow” threshold in which species can spend their life in a habitat favorable to their needs (chiussi et al., 2009). finally, microhabitat simulation using the phabsim software was used to estimate eflow requirements for salmonids and brown trout in the cordevole river basin. this study case demonstrated that a fixed water release of 600 l s-1 could be sufficient for the maintenance of good quality habitat for salmons, whereas a water releases of at least 350 l s-1 is sufficient for the maintenance of a good quality condition for brown trout’s habitat (schmidt and benítez sanz, 2012). more recently, other models have used alternative approaches based on a mix of expert opinions and fuzzy logic-based rules to describe the habitat use by target species and for the application of hydraulic-habitat models. a modified ifim micro-habitat method, using a fuzzy logic-based model, was applied to the barbel (barbus plebejus) and the chub (leuciscus cephalus) habitat suitability response to hydrologic and water quality parameters variation in the arno and serchio rivers. this trial provided evidence that fuzzy logicbased estimates of the target species’ response to critical environmental variations are more reliable than those obtained through the conventional ifim approach (marsili-libelli et al., 2013). e-flow implementation in sardinia (italy) sardinia is the second largest island in the mediterranean sea. it is characterized by a typically mediterranean climate with warm and moderately rainy winters and dry and hot summers. sardinian rivers are characterized by a strong variability of the water outflow, due to seasonal hydrological oscillations (gasith and resh 1999; de waele et al., 2010), and they constitute the predominant water contribution to the island water resources since the contributions from snowmelt and groundwater are almost negligible. in the last decades, a drastic mean yearly runoff decrease (up to ca 40% over the 1975-2010 period when compared with the previous 50 years; montaldo and sirigu, 2017) has been registered for sardinian basins. this caused relevant water scarcity issues and led to the construction of more than 50 dams, to store water in critical periods of the year. winter precipitations will reduce in the near future because of expected changes in the north atlantic oscillation (nao) index associated with climate change (montaldo and sirigu, 2017). this will likely determine a further decrease of water resources availability in sardinia, which, in turn, will require an increase in the number of hydro-morphologic alterations to satisfy the water demand for agriculture and human uses. additionally, this scenario will also be exacerbated because of unpredictable extreme rainfalls, to which sardinia is being more and more vulnerable. for example, in years from 2004 to 2018, these events became particularly frequent (at least once per year) and often produced flash floods causing severe changes in channel morphology and sediment distribution, damage to private and public infrastructure, and even human deaths (bodini et al., 2010; de waele et al., 2010). given the present and future critical scenario in rivers’ runoff and their predictable strong impacts on freshwater habitats, new strategies and designs in water resources planning and management are required. this need appears crucial, especially for the assessment of a sustainable eflow regulation plan aimed at ensuring the maintenance of key (including endemic) habitats and species. to date, the sardinian water basin management plan simply requires the maintenance of a fixed mvf equal to the 10% of the natural flow regime (i.e., the natural water flow, excluding any withdraw or artificial input). however, in 2018, following the dd 30/sta, the regional ecological flows resolution n.8/2018. this directive sets out a methodological approach for the determination of the e-flow in order to ensure the maintenance, in watercourses, of the ecological flow to support the achievement of environmental objectives defined in accordance with directive 2000/60/ec. this directive updated the formula to estimate e-flow as follows: no nco mm er cia l u se on ly implementation of the eu ecological flow policy in italy with a focus on sardinia 29 eflow=qn (t,f)× k where qn(t,f) is the water flow of the watercourse at its natural condition in the time t as a function of the monthly average rainfall f, and k is a coefficient calculated from the combined assessment of the current river ecological status, the analysis of the pressures affecting the river, and the hydro-morphological and biological parameters. according to the directive n.8/2018, this new formula, and its hydro-morphological and biological parameters, has to be tested and validated with specific technicalscientific experimentation for the determination of site-specific e-flow values. since none experimental application of hydraulic-habitat models have been tested in the sardinian territory, this formula has been identified as the basis for a rapid, homogeneous district-scale e-flow implementation plan, coping with the large heterogeneity of eco-hydraulic conditions of the sardinian rivers. in particular, the adoption in the formula of a rainfall-related coefficient, of patterns in flow regime at either intraand inter-annual temporal scales and of correction coefficients considering hydro-morphological and biological targets is expected to allow the development of a more reliable and ges-oriented e-flow assessment, integrated into the wfd process. challenges and limitations e-flow science has evolved substantially in the last decades, with a growing number of experts involved to develop new approaches and methods to respond to the numerous challenges that the implementations of the eflow policy triggers (acreman, 2016). however, despite the increasing number of published e-flow related studies, approaches, models, methods, and experimental applications to numerous locations, remain a global lack of achievement of appropriate e-flow assessment policy, mainly due to the conflicts arising between the preservation of the riverine environment and the management and exploitation of water resource for human needs (le quesne et al., 2010). in europe, considerable progress has been made on the application of the e-flow through the implementation of the water framework directive (european commission, 2015). the cis guidance document no. 31 provided an important and useful tool for mss to refine e-flow targets by developing new methods, spanning from relatively simple hydrologic to the highly sophisticated ones. however, significant challenges remain to achieve the ambitious objectives of the wfd, especially with regard to the e-flow assessment. some of these challenges revolve around the integration of macro-scale considerations into regional and basin-scale integrated water management plans. in a context of changing society, increasing population and changing climate, the water flow governance has to be adapted to the territorial specificities, in terms of regulatory and institutional frameworks, cultural practices, climatic, geographical and economic conditions. in italy, the e-flow matter has been faced in a more concrete way only in recent years. the introduction of the dd 30 / sta in 2018 imposed to all the district authorities a substantial change in the water management policies, in terms of the amount of outflow released downstream dams and weirs and the release of new water withdrawal concessions. this process led to move from a merely water management oriented mvf concept to another one oriented primarily at preserving the river ecosystems and maintain them in a good environmental status. a more detailed approach is needed, as suggested by the dd 30/sta, to update the already available methods with new ones. a spatial downscaling process of e-flow assessment will require new quantitative and qualitative approaches involving a multidisciplinary panel of experts (e.g., engineers, hydrologists, hydraulic modellers, ecologists and socio-economists) but also a large effort to test, by an accurate monitoring plan, the reliability and efficacy of the estimated e-flow values in the field. ideally, the process should become so accurate and reliable to allow a feedback monitoring approach, by which corrections to the e-flow will be adopted pending the availability of early warning indicators of environmental change. this approach, by cascade, will allow the design and operability of management measures able to correct any gap from the expected ges. this process, again, will need an accurate implementation of e-flow values and monitoring of its efficacy at the different spatial scales, from the regional to the basin one. this will require a huge effort in ameliorating the knowledge of biological and ecological processes sustaining the ges at the different spatial scales, which will require more and more commitments to either basic or applied research. this, in our view, will hold true especially for the case of sardinia, where the management of water resources is heavily dependent on the socio-economic and climate contexts. in a territory that has been suffering for the past ten years one of the most critical water scarcity conditions, numerous technical and logistical challenges will presumably arise. for example, the necessity of providing the needed water for agriculture, industrial or human usages insight of critical periods will likely push the authorities to never allow the release of a proper eflow. the need of different approaches to be adopted for very different rivers arises if we consider the plausible difficulties in establishing an e-flow value for intermittent or even ephemeral rivers (as the majority of those encountered in sardinia) when compared with the no nco mm er cia l u se on ly d. moccia et al.30 perennial ones (theodoropoulos et al., 2019). a critical issue for the application of e-flow, wherever defined and quantified, remains peculiarly for the sardinian territory the limited infrastructural tools to manage waters’ release from dams built in a time when e-flow was not considered at all and, therefore, which are not equipped, yet, with the technology able to operate in a reasonable time the needed water release in response to a change or decline in river ecosystem ecological integrity. arguably, in the future, a scientifically grounded implementation of water management infrastructures embodying the e-flow concept will be needed in order to address also the potential societal implication of water management, especially where this will generate conflicts between economic and ecological sustainability targets. in this context, hydraulic-habitat models can provide ecologically important data to calibrate hydrological methods and update e-flow release at different spatial and time scale within the italian territory. in italy, microhabitat approaches have been used mostly to correct minimum vital flow adopted in the river basin management plans. these methodologies tailored e-flow assessment at the site-specific scale, mostly downstream dams and hydropower stations, for only a few fish species and their vital stages. this approach can be useful when there is the necessity of focusing on one endangered species or when dealing with a specific e-flow release situation. nevertheless, due to their dependence on hydrological models, they are unreliable for low flow and non-flow periods. these limitations make this approach not well suitable for the definition of a regional or basinscale e-flow assessment, especially in a water-scarce region such as sardinia. differently, the application of meso-scale methodologies, in particular the mesohabsim method, results to be more suitable to the italian territory as it can be adequately integrated into a multi-criteria analysis framework allowing the calculation of e-flow criteria from site-specific to the regional and basin-scale (vezza et al., 2014; parasiewicz et al., 2019). furthermore, using deductive models, based on conditional habitat suitability criteria derived from expert knowledge and literature-based information, mesohabsim model can provide general information about habitat suitability even with a small amount of field data. another important benefit of this approach, especially in a heterogeneous territory such as italy, is the applicability of this model to different hydromorphological river types, allowing to describing complex habitat dynamics also in temporary rivers (acuña et al., 2019) or where characteristics of the riverbed do not allow the use of hydraulic simulation models (vassoney et al., 2019). in conclusion, we suggest that the meso-habitat approach result particularly appropriate for modelling and evaluating the integrity of rivers’ habitats within the italian territory, with the aims of defining the regional and basin e-flow assessment. it could be also used in monitoring e-flow release from catchments and dams and rivers requalification projects, as well as evaluating the impact of hydro-morphological pressure over fish fauna of watercourses and conserving particular protected aquatic species. it is also worth to notice that the mesohabsim method uses an open-source software, which is reliable and traceable, and it expresses a judgement, through the habitat integrity index, which is divided into five classes, as for other indicators required by the european legislation. for the italian territory, the set of models available for the mesohabsim methodology includes 24 fish species (and relative vital stages), typical of the italian hydrographic system (simstream software, referred to in mlg ispra 154/2017). however, this method has been applied mostly on northern regions, in order to further developed and improved the integration of this method for e-flow assessment within the italian territory, is essential to conduct further field studies at local and regional scale to calibrate and validate habitat suitability index in different contexts and for different species. in particular, in the sardinian region, insight of the high climate change-driven flow alterations predicted for the future scenario, the best science-based information of species response to habitat limitation will be required. we suggest that mesohabsim method should be implemented together with the validation of the recently proposed formula to obtain a solid and reliable e-flow assessment methodology. today, more than ever, human actions in river ecosystems have become one of the primary contributions to the decline of freshwater habitats and species (postel and richter, 2003). worldwide, maintaining ecosystem functions and services at current levels in a warmer climate and under changing socioeconomic conditions is a major challenge for water resources managers. reducing pressures on water quality and maintenance of natural flow regimes are two of the main management actions that can help to ensure habitats conservation, at the same time favouring species resilience to flow alterations (dudgeon et al., 2006). thus, we further conclude that the definition of handling and reliable e-flow assessment methods embedding the expected consequences of natural and anthropogenic disturbance with the need of nature conservation is to be considered a basic priority of any water management plan. acknowledgements this study has been carried out in the framework of a collaboration between the head office of the regional agency of the sardinian river basin district (servizio tutela e gestione delle risorse idriche, vigilanza sui no nco mm er cia l u se on ly implementation of the eu ecological flow policy in italy with a focus on sardinia 31 servizi idrici e gestione delle siccità) and the departments of civil-environmental engineering and architecture (dicaar) and of life and environmental sciences (disva) of the university of cagliari, italy. corresponding author: mocciadavide@unica.it key words: ecological flow; minimum vital flow; water framework directive; anthropogenic pressures; italy; sardinia. received: 23 dicember 2019. accepted: 12 may 2020. this work is licensed under a creative commons attribution noncommercial 4.0 license (cc by-nc 4.0). ©copyright: the author(s), 2020 licensee pagepress, italy advances in oceanography and limnology, 2020; 11:8781 doi: 10.4081/aiol.2020.8781 references acreman m, 2016. environmental flows-basics for novices. wires water 3:622-628. acreman mc, dunbar mj, 2004. defining environmental river flow requirements a review. hydrol. earth syst. sci. 8: 861-876. acreman m, ferguson a, 2010. environmental flows and the european water framework directive. freshwater biol. 55:32-48. acuña v, jorda-capdevila d, vezza p, de girolamo a.m, mcclain me, stubbington r, pastor av, lamouroux n, von schiller d, munné a, datry t, 2020. accounting for flow intermittency in environmental flows design. j. appl. ecol. 57:742-753. adamczyk m, parasiewicz, p, vezza p, prus p, de cesare g, 2019. empirical validation of mesohabsim models developed with different habitat suitability criteria for bullhead cottus gobio l. as an indicator species. water 11:726. alfredsen k, tesaker e, 2002. winter habitat assessment strategies and incorporation of winter habitat in the norwegian habitat assessment tools. hydrol. process. 16:927-936. arthington ah, bhaduri a, bunn se, jackson se, tharme re, tickner d, young b, acreman m, baker n, capon s, horne ac, kendy e, mcclain me, poff nl, richter bd, ward s, 2018. the brisbane declaration and global action agenda on environmental flows (2018). front. envirn. sci. 6:45. arthington ah, tharme r, brizga so, pusey bj, kennard mj, 2004. environmental flow assessment with emphasis on holistic methodologies, p. 37-65. in: r. welcomm and t. petr (eds.). proceedings of the second international symposium on the management of large rivers for fisheries volume ii rap publication 2004/17, fao regional office for asia and the pacific, bangkok. arthington ah, zalucki jm, 1998. comparative evaluation of environmental flow assessment techniques: review of methods. land and water resources research and development corporation, canberra. barnett tp, adam jc, lettenmaier dp, 2005. potential impacts of a warming climate on water availability in snowdominated regions. nature 438:303-309. blöschl g, hall j, parajka j, perdigão rap, merz b, arheimer b, aronica t, chirico gb, bilibashi a, bonacci o, borga m, čanjevac i, castellarin a, chirico gb, claps p, fiala k, frolova n, gorbachova l, güll a, hannaford j, harrigan s, kireeva m, kiss a, kjeldsen tr, kohnová s, koskela jj, ledvinka o, macdonald n, mavrova-guirguinova m, mediero l, merz r, molnar p, montanari a, murphy c, osuch m, ovcharuk v, radevski i, rogger m, salinas jl, sauquet r, šraj m, szolgay j, viglione a, volpi e, wilson d, zaimi k, živković n, 2017. changing climate shifts timing of european floods. science 357:588-590. bodini a, cossu qa, 2010. vulnerability assessment of centraleast sardinia (italy) to extreme rainfall events. nat. hazard. heart. sys. 10: 61-72. bovee kd, 1982. a guide to stream habitat analysis using the ifim. us fish and wildlife service report fws/obs-82/ 26. fort collins, usa. bovee kd, lamb bl, bartholow jm, stalnaker cb, taylor j, henriksen j, 1998. stream habitat analysis using the instream flow incremental methodology. information and technology report usgs/brd/itr-1998-0004. fort collins, co: u.s. geological survey-brd. brunetti m, colacino m, maugeri m, nanni t, 2001. trends in daily intensity of precipitation in italy from 1951 to 1996. int. j. climatol. 21:299-316. bunn se, arthington ah, 2002. basic principles and ecological consequences of altered flow regimes for aquatic biodiversity. environ. manage. 30:492-507. caissie d, el-jabi, 1995. comparison and regionalization of hydrologically based instream flow techniques in atlantic canada. can. j. civ. eng. 22:235-246. chiussi s, bodini a, bondavalli c, pattini l, 2009. [il caso del fiume taro (provincia di parma)].[article in italian]. biologia ambientale 23:56-65. coppola e, giorgi f, 2010. an assessment of temperature and precipitation change projections over italy from recent global and regional climate model simulations. int. j. climatol. 30:11–32. de girolamo am, bouraoui f, buffagni a, pappagallo g, lo porto a, 2017. hydrology under climate change in a temporary river system: potential impact on water balance and flow regime. river. res. app. 33:1219-1232. de waele j, martina mlv, sanna l, cabras s, cossu qa, 2010. flash flood hydrology in karstic terrain: flumineddu canyon, central-east sardinia. geomorphology 120:162–173. döll p, fiedler k, zhang j, 2009. global-scale analysis of river flow alterations due to water withdrawals and reservoirs. hydrol. earth syst. sc. 13:2413-2432. dudgeon d, arthington ah, gessner mo, kawabata zi, knowler dj, lévêque c, naiman r, prieur-richard ah, soto d, stiassny mlj, sullivan ca, 2006. freshwater biodiversity: importance, threats, status and conservation challenges. biol. rev. camb. philos. soc. 81:163-82. dyson m, bergkamp m, scanlon j, 2003. flow: the essentials of environmental flows. iucn, gland. no nco mm er cia l u se on ly d. moccia et al.32 e.e. agency, report n.7/2018. european waters assessment of status and pressures 2018. efstratiadis a, tegos a, varveris a, koutsoyiannis d, 2014. assessment of environmental flows under limited data availability: case study of the acheloos river, greece. hydrolog. sci. j. 5:731-750. european commission, 2015. ecological flows in the implementation of the water framework directive. wfd cis guidance document no. 31. feyen l, dankers r, 2009. impact of global warming on streamflow drought in europe. j. geophys. res. atmos. 114:1-17. filipe af, lawrence je, bonada n, 2013. vulnerability of stream biota to climate change in mediterranean climate regions: a synthesis of ecological responses and conservation challenges. hydrobiologia 719:331-351. fornaroli r, cabrini r, sartori l, marazzi f, canobbio s, mezzanotte v, 2016. optimal flow for brown trout: habitat – prey optimization. sci. total environ. 566-567:1568-1578. forslund a, renof̈al̈t bm, maijer k, krchnak k, cross k, smith m, mcclain m, davidson s, barchiesi s, farrell t, 2009. securing water for ecosystems and human well-being: the importance of environmental flows. swedish water house, stockholm. galland j, 1991. telemac: a new numerical model for solving shallow water equations. adv. water resour. 14:138-148. gasith a, resh vh, 1999. streams in mediterranean climate regions: abiotic influences and biotic responses to predictable seasonal events. annu. rev. ecol. syst. 30:51-81 giorgi f, 2006. climate change hot-spots. geophys. res. lett. 33:l08707. giorgi f, lionello p, 2008. climate change projections for the mediterranean region. global planet. change 63:90-104. gosselin mp, maddock i, petts g, 2012. mesohabitat use by brown trout (salmo trutta) in a small groundwaterdominated stream. river res. appl. 28:390-401. hauer c, mandlburger g, habersack h, 2009. hydraulically related hydro-morphological units: description based on a new conceptual mesohabitat evaluation model (mem) using lidar data as geometric input. river res. appl. 25:29-47. horne ac, o’donnell el, acreman m, mcclain me, poff nl, webb ja, stewardson mj, bond nr, richter b, arthington ah, tharme re, garrick de, daniell ka, conallin jc, thomas ga, hart bt, 2017. moving forward: the implementation challenge for environmental water management, p. 649-673. in: a.c. horne, j.a. webb, m. j. stewardson, b. richter and m. acreman (eds.), water for the environment: from policy and science to implementation and management. academic press, london. jowett i, 2010. rhyhabsim-river hydraulic and habitat simulation software. tech. rep. manual version 5.0. niwa. international commission on large dams 2007. world register of dams, paris. available from: www.icold-cigb.net king jm, brown c, 2006. environmental flows: striking the balance between development and resource protection. ecol. soc. 11:26. king jm, brown ca, sabet h, 2003. a scenario-based holistic approach for environmental flow assessments. river res. appl. 19:619-639. king jm, louw md, 1998. instreamflow assessments for regulated rivers in south africa using the building block methodology. aquat. ecosyst. health manag. 1:109-124. king jm, tharme re, de villiers ms, 2008. environmental flow assessments for rivers: manual for the building block methodology. wrc report no tt 354/08. cape town:364. lake p, 2007. flow-regulated disturbances and ecological response: floods and droughts. in: p.j. wood, d.m. hannah and j.p. sadler (eds.), hydroecology and eco-hydrology: past, present and future. j. wiley & sons ltd., chichester. le quesne t, kendy e, weston d, 2010. the implementation challenge. taking stock of government policies to protect and restore environmental flows. wwf (world wide fund for nature) and tnc (the nature conservancy). lionello p, abrantes f, congedi l, dulac f, gacic m, gomis d, goodess c, hoff h, kutiel h, luterbacher j, planton s, reale m, schröder k, struglia mv, toreti a, tsimplis m, ulbrich u, xoplaki e, 2012. introduction: mediterranean climate: background information, p. xxxv-xc. in: p. lionello (ed.), the climate of the mediterranean region. from the past to the future. elsevier. doi: 10.1016/b978-0-12-416042-2.00012-4. maddock i, bickerton m, spence r, pickering t, 2001. reallocation of compensation releases to restore river flows and improve instream habitat availability in the upper derwent catchment, derbyshire, uk. regul. rivers-res. manage. 17:417-441. magilligan fj, nislow kh, 2005. changes in hydrologic regime by dams. geomorphology 71:61-78. marsili-libelli s, giusti e, nocita a, 2013. a new instream flow assessment method based on fuzzy habitat suitability and large-scale river modelling. environ. modell. softw. 41:27–38. montaldo n, sarigu a, 2017. potential links between the north atlantic oscillation and decreasing precipitation and runoff on a mediterranean area. j. hydrol. 553:419-437. nilsson c, reidy ca, dynesius m, revenga c, 2005. fragmentation and flow regulation of the world’s large river systems. science 308:405-408. oecd, 2012. oecd environmental outlook to 2050: the consequences of inaction. organisation for economic cooperation and development, paris. parasiewicz p, 2001. mesohabsim: a concept for application of instream flow models in river restoration planning. fish. res. 26:6-13. parasiewicz p, 2007. the mesohabsim model revisited. river res. appl. 23:893-903. parasiewicz p, rogers jn, vezza p, gortazar j, seager t, pegg m, wiśniewolski w, comoglio c, 2013. applications of the mesohabsim simulation model, p. 109-124. in: i. maddock, p. kemp and p. wood (eds.), ecohydraulics: an integrated approach. j. wiley & sons ltd., chichester. parasiewicz p, prus, p, theodoropoulos, c, alfredsen k, adamczyk m, comoglio c, vezza p, 2019. editorial: environmental flows determination and monitoring with hydraulic habitat models pushing the boundaries of habitat models application. water 11:1950. peñas fj, barquín j, álvarez c, 2016. assessing hydrologic alteration: evaluation of different alternatives according to data availability. ecol. indic. 60:470-482. pletterbauer f, melcher a, graf w, 2018. climate change no nco mm er cia l u se on ly implementation of the eu ecological flow policy in italy with a focus on sardinia 33 impacts in riverine ecosystems, p. 203-223. in: s. schmutz and j. sendzimir (eds.), riverine ecosystem management. springer, cham. poff nl, allan jd, bain mb, karr jr, prestegaard kl, richter bd, sparks re, stromberg jc, 1997. the natural flow regime. bioscience 47:769-784 poff nl, matthews jh, 2013. environmental flows in the anthropocence: past progress and future prospects. curr. opin. env. sust. 5:667-675. poff nl, richter bd, arthington ah, bunn se, naiman rj, kendy e, acreman m, apse c, bledsoe bp, freeman mc, henriksen j, jacobson rb, kennen jg, merritt dm, o’keeffe jh, olden jd, rogers k, tharme re, warner a, 2010. the ecological limits of hydrologic alteration (eloha): a new framework for developing regional environmental flow standards. freshwater biol. 55:147-170. poff nl, zimmerman jkh, 2010. ecological responses to altered flow regimes: a literature review to inform the science and management of environmental flows. freshwater biol. 55:194-205. postel sl, 1998. water for food production: will there be enough in 2015. bioscience 48:629-637. postel s, richter bd, 2003. rivers for life: managing water for people and nature, ecological economics. island press, washington, dc. richter bd, 2009. rethinking environmental flows: from allocations and reserves to sustainability boundaries. river res. appl. 26:1052-1063. richter bd, warner at, meyer jl, lutz k, 2006. a collaborative and adaptive process for developing environmental flow recommendations. river res. appl. 22:297-318. salmoiraghi g, morolli c, 2007. [aggiornamento ed integrazione attività di studio per la determinazione sperimentale dei valori di deflusso minimo vitale (dmv) per il fiume marecchia].[report in italian]. autorità di bacino interregionale marecchia-conca: 74 pp. schmidt g, benítez sanz c, 2012. analysis of the implementation of environmental flows in the wider context of the river basin management plans. technical report. available from: https://ec.europa.eu/environment/archives/water/implrep2007 /pdf/water%20abstraction%20and%20use%20-%20eflows. pdf schneider m, jorde k, zoellner f, kerle f, 2001. development of a user-friendly software for ecological investigations on river systems, integration of a fuzzy rulebased approach, p. 354-360. in: l.m. hilty and p.w. gilgen (eds.), sustainability in the information society. metropolis, marburg. schneider m, noack m, gebler t, kopecki i, 2010. handbook for the habitat simulation model casimir-fish. base version. spinoni j, vogt jv, naumann g, barbosa p, dosio a, 2017. will drought events become more frequent and severe in europe? int. j. climatol. 381718-1736. steffler p, blackburn j, 2002. river 2d: two-dimensional depth averaged model of river hydrodynamics and fish habitat. introduction to depth averaged modelling and user’s manual. university of alberta. surian n, rinaldi m, 2003. morphological response to river engineering and management in alluvial channels in italy. geomorphology 50:307-326. tennant dl, 1976. instream flow regimens for fish, wildlife, recreation and related environmental resources. fisheries 1:6-10. tharme re, 1996. review of international methodologies for the quantification of the instream flow requirements of rivers. water law review final report for policy development for the department of water affairs and forestry, pretoria. freshwater research unit, university of cape town. tharme re, 2000. an overview of environmental flow methodologies, with particular reference to south africa, p. 15-40. in: j.m. king, r.e. tharme and m.s. de villiers (eds.), environmental flow assessments for rivers: manual for the building block methodology. water research commission technology transfer report no. tt131/00. water research commission: pretoria. tharme re, 2003. a global perspective on environmental flow assessment: emerging trends in the development and application of environmental flow methodologies for rivers. river res. appl. 19:397-441. the brisbane declaration, 2007. environmental flows are essential for freshwater ecosystem health and human wellbeing. proceedings 10th int. river symp. and int. environ. flows conf., 2007, brisbane. available from: https://www.conservationgateway.org/conservationpractice s/freshwater/environmentalflows/methodsandtools/elo ha/pages/brisbane-declaration.aspx theodoropoulos c, papadaki c, vardakas l, dimitriou e, kalogianni e, skoulikidis n, 2019. conceptualization and pilot application of a model-based environmental flow assessment adapted for intermittent rivers. aquat. sci. 81:10. tomozeiu r, pasqui m, quaresima s, 2017. future changes of air temperature over italian agricultural areas: a statistical downscaling technique applied to 2021–2050 and 2071– 2100 periods. meteorol. atmos. phys. 1-21. tonkin jd, poff nl, bond nr, horne a, merritt dm. reynolds lv, ruhi a, lytle da, 2019. prepare river ecosystems for an uncertain future. nature 570:301-303. u.s. geological survey (usgs), 2001. phabsim user’s manual. vassoney e, mammoliti mochet a, rocco r, maddalena r, vezza p, comoglio c, 2019. integrating meso-scale habitat modeling in the multicriteria analysis (mca) process for the assessment of hydropower sustainability. water 11:640. vezza p, ghia d, fea g, 2016. quantitative habitat models for the conservation of the endangered european crayfish austropotamobius pallipes complex (astacoidea: astacidae). in: t. kawai and n. cumberlidge (eds.), a global overview of the conservation of freshwater decapod crustaceans. springer, cham. vezza p, parasiewicz p, calles o, spairani m, comoglio c, 2014a. modelling habitat requirements of bullhead (cottus gobio) in alpine streams. aquat. sci. 76:1-15. vezza p, parasiewicz p, rosso m, comoglio c, 2012. defining minimum environmental flows at regional scale: application of mesoscale habitat models and catchments classification. river res. applic. 28:717-730. vezza p, parasiewicz p, spairani m, comoglio c, 2014b. habitat modelling in high-gradient streams: the mesoscale approach and application. ecol. appl. 24:844-861. viganò g, confortola g, fornaroli r, cabrini r, canobbio s, no nco mm er cia l u se on ly d. moccia et al.34 mezzanotte v, bocchiola d, 2016. effects of future climate change on a river habitat in an italian alpine catchment. j. hydrol. eng. 21:1-14. vismara r, azzellino a, bosi r, crosa g, gentili g, 2001. habitat suitability curves for brown trout (salmo trutta fario l.) in the river adda, northern italy: comparing univariate and multivariate approaches. regul. river. 17:37-50. wilkes ma, maddock i, link o, habit e, 2015. a communitylevel, mesoscale analysis of fish assemblage structure in shoreline habitats of a large river using multivariate regression trees. river res. appl. 32:652-665. zang cf, liu j, van der velde m, kraxner f, 2012. assessment of spatial and temporal patterns of green and blue water flows under natural conditions in inland river basins in northwest china. hydrol. earth syst. sci. 16:2859-2870. zhang q, gu x, singh vp, chen x, 2015. evaluation of ecological instream flow using multiple ecological indicators with consideration of hydrological alterations. j. hydrol. 529:711-722. no nco mm er cia l u se on ly layout 1 holothurian’s exploitation the worldwide consumption of fish food products from 1961 to 2017 increased at an average annual rate of 3.1%, with a consumption per capita of fish food raising from 9.0 kg to 20.5 kg in the same period (fao, 2020). although the wild caches have been followed by the development of fish farming, the state of the wild stocks has continued to decline with less than 66% of the stocks harvested in a sustainable way (fao, 2020). in response to the over-exploitation of wild finfish stocks, the invertebrate fisheries rapidly increased, being a new available source of seafood proteins and socio-economic opportunities (berkes et al., 2006, anderson et al., 2008). many of the new target species now belong to low trophic levels, as a response to the overall down effect of trophic webs caused by top predators (pauly et al., 2002; anderson et al., 2008). in many cases, the pressure on stocks within low trophic levels increased faster than their management policies (anderson et al., 2011a, 2011b), causing the spread of unregulated fishery and raising concerns for the possible consequences on ecosystem functioning and the sustainability of the fishery (andrew et al., 2002; leiva and castilla 2002; berkes et al., 2006; anderson et al., 2008; fao, 2008). sea cucumbers, marine invertebrates belonging to the echinodermata phylum, include more than 1500 species (horton et al., 2018) and, mainly being deposit feeders, represent a good example of low trophic level organisms. their fisheries had rapidly grown and expanded since 1980 as a consequence of the increasing demand from international markets, aquaculture and biomedical research programs (bordbar et al., 2011). holothurians are present in almost all the marine biotopes, from the littoral to hadal depths (purcell et al., 2012). holothurians are part of the chinese culinary tradition, are considered gourmet and luxury seafood and are generally sold as a dried product called bêche-de-mer or trepang (wen et al., 2010; yang and bai, 2015). the market price of this product depends on the quality (grade low, medium, high) (ram et al., 2014), with some particularly valuable species as apostichopus japonicus selenka, which holds the highest price of 2950 us$ dried kg−1, followed by holothuria scabra jaeger, 1833 (115640 us$ dry kg−1), holothuria lessoni massin, uthicke, purcell, rowe & samyn, 2009 (240-790 us$ dry kg−1) (purcell et al., 2012). the presence of high-value nutrients such as vitamin a, vitamin b1 (thiamine), vitamin b2 (riboflavin), vitamin b3 (niacin), and minerals (i.e., calcium, magnesium, iron and zinc) indicate that sea cucumbers are suitable tonic and restorative products, also rich in crude proteins (range 41-63%) (wen et al., 2010, bordbar et al., 2011). moreover, sea cucumbers, containing a number of biological and pharmacological bioactive compounds, have attracted attention for their potential medical value (bordbar et al., 2011). sea cucumbers contain numerous bioactive and anti-age substances that are already exploited in the cosmetic and pharmaceutical industries (fredalina et al., 1999; saito et al., 2002; zhao et al., 2007; bordbar et al., 2011; purcell, 2014). all these properties and the high market price led to the overexploitation and decline of sea cucumbers indo-pacific populations and the expansion of the fishery to reach new virgin stocks in galapagos islands, mexico, north america and the mediterranean sea (conand, 2006; purcell et al., 2012; gonzàlez-wangüemert et al., 2018). the estimated sea cucumbers harvest, from asia and pacific regions, ranges from 20.000 to 40.000 t per year of the dry product (fao, 2012). fisheries from african and indian ocean regions also contribute to the complex amount with the range of 2000-2500 t per year (fao, 2012). review biology, ecology and management perspectives of overexploited deposit-feeders sea cucumbers, with focus on holothuria tubulosa (gmelin, 1788) viviana pasquini,*1 ambra angelica giglioli,1 antonio pusceddu,1 pierantonio addis1 1department of life and environmental sciences, university of cagliari, via t. fiorelli 1, 09126 cagliari, italy abstract the increasing harvesting of low trophic level organisms is raising concern about the possible consequences on the ecosystem functioning. in particular, the continuous demand of sea cucumbers from the international market led to the overexploitation of either traditionally harvested or new target species, including the mediterranean ones. sea cucumbers are mostly deposit feeders able to consume sedimentary organic matter and, thus, are ideal candidate for the remediation of eutrophicated sediments, like those beneath aquaculture projects. breeding and restocking of overexploited sea cucumbers populations are well-established practices for indo-pacific species like holothuria scabra and apostichopus japonicus. some attempts have also been made for the mediterranean species holothuria tubulosa, but, so far, the adaptation of protocols used for other species has presented several issues. we here summarize narratively the available information about sea cucumbers rearing protocols with the aim of identifying their major flaws and gaps of knowledge and fostering research about new triggers for spawning and feasible protocols to reduce the high mortality of post-settlers. no nco mm er cia l u se on ly biology, ecology and management of deposit-feeders sea cucumbers 37 less information are available about sea cucumbers fisheries in the mediterranean sea, in particular for holothuria tubulosa gmelin, 1788, holothuria mammata grube, 1840, holothuria sanctori delle chiaje, 1823, holothuria forskali delle chiaje, 1823, parastichopus regalis cuvier, 1817, and holothuria arguinensis koehler & vaney, 1906 (çakly et al., 2004; antoniadou and vafidis, 2011; sicuro and levine 2011; gonzàlez-wangüemert and borrero-perez, 2012; mezali and thandar 2014; gonzalez-wangüemert et al., 2014a, 2015). presently, more than half of global sea cucumber fisheries are considered depleted or overexploited to the extent that governments (including the italian government) have banned their harvesting (anderson et al., 2011; gonzález-wangüemert et al., 2014, 2018). with the 38% of sea cucumber fisheries currently unregulated and an unknown level of illegal catches, this fishery is considered unsustainable and far from being adequately managed (anderson et al., 2011; choo, 2008; toral-granda, 2008). the unregulated exploitation of sea cucumbers is a rising concern for their conservation, with 16 species worldwide now classified as “vulnerable” or “endangered”, according to the iucn red list (conand et al., 2014, ramírez-gonzález et al., 2020). concern also raises because most of the harvested sea cucumbers are depositfeeders, thus playing an ecological key role due to their feeding behaviour (uthicke, 2001; roberts et al., 2000), their decline could have severe consequences on sedimentary biogeochemistry and benthic ecosystem functioning. here we reviewed the available information about the ecological role of sea cucumbers, with a focus on the mediterranean h. tubulosa, their breeding, fishery management issues, main gaps of knowledge and future perspectives for their use as remediation of eutrophicated sediments. life history and population dynamics of sea cucumbers the increasing interest towards sea cucumbers and their use for food, medical and habitat remediation purposes, stimulated exploration about their reproductive cycle and population dynamics, both crucial aspects for the assessment of wild stocks and their eventual management. almost all sea cucumbers are broadcast spawners with external fertilization that present an annual or bi-annual maturation season (mercier and hamel, 2009; mohsen and yang, 2021). with a few exceptions of hermaphrodite species, they are generally gonochoric that leak in sexual dimorphisms (smiley et al., 1991; mercier and hamel, 2009). the life cycle of sea cucumbers is characterised by one or more planktonic larval stages starting with a feeding auricularia (early, mid and late), a non-feeding doliolaria and then a feeding pentactula that settle on the substrate (strathmann, 1975; ito and kitamura, 1997; yanagisawa, 1998). doliolaria actively explore the surrounding environment to identify the best place to settle and made the last metamorphosis into the pentactula. if the conditions are not suitable for settlement, the larvae will keep swimming for several days (mercier et al., 2000). the pentactula lose the ability to swim but can continue to explore the surrounding environment with the buccal podia, moving by small jumps (mercier et al., 2000). although rarely, evans and palmer (2003) reported the ability of the pentactula larvae of parastichopus californicus stimpson, 1857, to clone, forming a bud that, after separation, will normally develop into an auricularia larvae. the pentactula larvae will start to feed and grow, becoming a juvenile in a variable time lag (mercier et al., 2000; agudo, 2006; mercier and hamel, 2009; rakaj et al., 2018, 2019). information about the mechanisms of settlement, physiology and cue that can stimulate the larvae to settle are poorly explored and understood, so far. studies conducted in mesocosm investigated the success of the larval settlement, which can strongly depend on the larval nutrition state and the capacity to accumulate lipids (peters-didier and sewell, 2019). in the late auricularia stage of h. scabra, the development of the hyaline spheres indicates an adequate feeding, and their size is a reliable indicator for subsequent performance (duy et al., 2016). the settlement and the last metamorphosis, as for other echinoderms, represents a survivorship bottleneck that can lead to high mortality rates. the early juvenile stage (<5 mm length) is also vulnerable and a critical phase with substantial mortality rates (agudo, 2006; rakaj et al., 2018). the holothurians recruitment has been studied mainly on historically exploited species, and information about post-settlers and juveniles in the field is scarcely recorded in the literature and, even, referred to sporadic occasions. for instance, the recruitment of h. scabra has been found to occur on a monthly time scale on seagrasses, with adult specimens mainly observed in sandy sediments and juveniles in organic matter (om) enriched muddy sediments (mercier et al., 2000). the lack of other information about holothurians recruitment can also be ascribed to the potential misidentification of the species because they can have a considerably different morphology when compared with that of adults. besides this, juveniles might occupy different habitats and can be obscured from the researchers’ view because of their cryptic behaviour (shiell, 2004). h. scabra juveniles can also be affected by predation-mediated mortality by fish belonging to the balistidae, labridae, lethrinidae and nemipteridae families (dance et al., 2003), sea stars, and crustaceans (kinch et al., 2008). holothurians’ recruitment can also be affected by geographic distances, the duration of the larval period and to no nco mm er cia l u se on ly v. pasquini et al.38 the hydrodynamic retainment in coastal areas (uthicke, et al., 1998, 1999, 2001; uthicke and purcell, 2004). most studies about holothurians’ population dynamics explored species with a long history of exploitation, including a. japonicus, cucumaria frondosa gunnerus, 1767, and isostichopus fuscus ludwig, 1875, (herreropérezrul et al., 1999; reyes-bonilla and herrero-pérezrul, 2003; hamel and mercier, 2008; anderson et al., 2011; purcell et al., 2011; yang et al., 2015; glockner-fagetti et al., 2016). unfortunately, the absence of a rigid structure in sea cucumbers and the high plasticity of the body wall make it difficult to investigate the growth rates of holothurians. alternative methods proposed include marking the calcareous (epi-pharyngeal) ring, chemical marking of spicules, external and internal tagging (kinch et al., 2008). however, all of these methods are affected by wide methodological biases but also by the bio-ecological traits of holothurians. in fact, the body size of holothurians can vary as a response to changing environmental conditions (tolon et al., 2017b), the occurrence of asexual reproduction through fission (purwati and dwiono, 2005; uthicke and conand, 2005; laxminarayana, 2006; purwati and dwiono, 2007; purcell et al., 2012; dolmatov, 2014,2021) or the evisceration of their internal organs (intestine, gonads and respiratory trees) through autotomy, in response to predation and other environmental stressors (shukalyuk and dolmatov, 2001; wilkie, 2001; spirina and dolmatov, 2003; zang et al., 2012). the evisceration is a typical behavioural trait of holothurians that does not lead to the death of the organism, rather is followed by the re-growth of the internal organs (dawbin, 1949; murray and garcía-arrarás, 2004; garcía-arrarás et al., 2006; dolmatov and ginanova, 2009). interestingly, after evisceration, the respiratory function shifts to the body wall for the time necessary for the respiratory trees’ regrowth. during this period, sea cucumbers will consume endogenous substances, which causes a significant body weight loss (zang et al., 2012, zhang et al., 2017). because of the multiple factors regulating holothurians body size, small individuals are not necessarily the youngest ones (kinch et al., 2008). breeding of sea cucumbers the development of sea cucumbers ex situ breeding protocols derived from the need to reduce the pressure on wild overexploited stocks. breeding sea cucumbers can be used for restocking activities (purcell and kirby, 2006) as already explored for other exploited echinoderms (couvray et al., 2015; giglioli et al., 2021). moreover, producing and releasing juveniles sea cucumbers reared in “conservation hatchery”, could be a useful tool for bioremediation of eutrophicated sediments or in integrated multi-trophic aquaculture systems (see below) without burden on wild populations. the experimental reproduction of sea cucumbers has been carried out for many species and the aquaculture is now established for largely exploited indo-pacific species like h. scabra (agudo, 2006) and a. japonicus (purcell et al., 2012; shi et al., 2013, 2015; pietrak et al., 2014). china, the largest consumer and producer country, is breeding annually about 10 000 t of dry weight a. japonicus from aquaculture to supply the local demand, while in other countries this activity is still in a pilot scale or in early development stages (choo, 2008). it has been estimated that once released in the field h. scabra can reach the commercial size of 700 g ind-1 in about 2-3 years, with a survivorship of 7-20% (purcell and simutoga, 2008). in the last decade, new attempts have been also made with the mediterranean species h. tubulosa and holothuria polii delle chiaje, 1823, (rakaj et al., 2018, 2019); h. arguinensis (domínguez-godino et al., 2015); h. mammata (domínguez-godino and gonzález-wangüemert, 2018). feeding behavior and ecological role of sea cucumbers deposit-feeders holothurians acquire food by swallowing large volumes of sediment (ramon et al., 2019). they sift through the sediment with tentacles and feed on detritus, organic matter, sand and the relative grown-over biofilm, expelling sandy pellets after digestion (hartati et al., 2020). the feeding starts with capturing the sedimentary food particles with tentacles and their release into the pharynx through the circum-oral tentacles. once inside the mouth the particles are mixed with the digestive enzymes and compressed into a plug which moves throughout the gut following a plug-flow reactor model. the plug is then transported by peristalsis along the simple digestive system that ends in the posterior part of the animal (zamora and jeffs, 2011). sea cucumbers predominantly feed on sedimentary organic detritus associated with micro-organisms and small benthic organisms (roberts et al., 2000). in the gut mineral and organic particles are found along with fragments of shell, barnacles, seagrasses, echinoderms ossicles, faecal pellets, foraminifera shells, with a highly variable size (roberts et al., 2000). information about the potential selectivity of shallowwater holothurians is controversial. some holothurians are able to choose om enriched particles, whereas others appear not to be (moriarty 1982; hammond, 1983; uthicke and karez, 1999; battaglene et al., 1999; slater et al., 2011; navarro et al., 2013; sun et al., 2015; lee et al., 2018; hartati et al., 2020). the selective ability can be related to how sea cucumbers feed on the sediment, which is highly variable among species, depending on their tenno nco mm er cia l u se on ly biology, ecology and management of deposit-feeders sea cucumbers 39 tacles dimension, the size and gut morphology (roberts et al. 2001, dar and ahmad, 2006; ramón et al., 2019). the selection of smaller organic-rich particles might be due to the greater ease of being caught and held by the tentacles, or to the potential chemo-selection ability of holothurians (schneider et al 2013; lee et al., 2018). the presence of a higher om content in the gut compared to the one present in the sediment can be a consequence of a passive selection of the finest grain size of the particles which can be more easily ingested. this, in turn, can be explained because smaller grain size particles can have a higher om content due to the wider surface available for the microbial colonization (hargrave, 1972; levinton, 1972; dale, 1974; yamamoto and lopez, 1985; manini and luna, 2003). considering their feeding behaviour, sea cucumbers are great seafloor bioturbators, able to rework large amounts of sediments via ingestion and excretion (9-82 kg ind−1 year−1) which can extensively blend and reform seafloor substrata (coulon and jangoux, 1993; uthicke and karez, 1999; mangion et al., 2004). bioturbation intensity can influence the sediment permeability, oxygen concentration, water content and chemical gradients in pore water, affecting the rate of remineralization and the inorganic nutrient flux and, finally, can redistribute food resources for the other benthos (reise, 2002; lohrer et al., 2004; solan et al., 2004; meysman, 2006a). bioturbation carried out by sea cucumbers can be circumscribed to the upper layer of the sediment or reach up to ten centimetres depth based on the habits of the species whether they are fossorial or not (uthicke and karez, 1999; purcell, 2004a; amaro et al., 2010). the role of holothurians in recycling the sedimentary om is considered one of their main ecosystem functions (purcell et al., 2016). the ability to reduce the om content in the sediment has been recently investigated (dar and ahmad, 2006; i̇şgören-emiroğlu and günay, 2007; slater and carton, 2009; wolkenhour et al., 2010; zamora and jeffs, 2011; tolon et al., 2017a; neofitou et al., 2019; hartati et al., 2020). the sea cucumber australostichopus mollis hutton, 1872, can significantly reduce total organic carbon (toc), chlorophyll-a and phaeopigments contents of sediments impacted by green-lipped mussel biodeposits (faeces and pseudofaeces) (slater and carton, 2009). mactavish et al. (2012) reported that a. mollis suppressed benthic microalgae and facilitated bacterial activity, causing a shift in the balance of benthic production and decomposition processes. juveniles of the same species decreased their ingestion rate with the increasing of the total sedimentary organic matter (tom), showing the ability of this species to use different amounts of tom, changing their feeding behaviour and digestive physiology (zamora and jeffs, 2011). h. tubulosa reduced the sedimentary om and organic carbon (oc) by 31-59%, with an absorption rate of 43 and 55% respectively, both in manipulative laboratory and field experiments (neofitou et al., 2019). the functioning of the digestive system of holothurians has been modelled and defined as a sort of ‘bioreactor’, where the ingested nutrients are quickly extracted and assimilated (penry and jumars, 1986, 1987; jumars, 2000; amaro et al., 2010). the grazing of holothurians could increase the exchange flux of nutrients across the sedimentwater interface and promote nutrient regeneration (zhou et al., 2006; yuan et al., 2013; slater and carton, 2009; slater et al., 2011; zamora and jeffs, 2011, 2012a, b). on the other hand, other species, like a. japonicus, could not affect toc and total nitrogen (tn) sedimentary contents, but can cause om particles redistribution and inhibit microphytobenthos (michio et al., 2003). the mediterranean sea cucumber holothuria tubulosa a new target species candidate for sea cucumbers aquaculture is holoturia tubulosa (gmelin 1788), one of the most common and widespread holothurians in the coastal areas of the mediterranean sea and the eastern atlantic ocean (tortonese, 1965; koukouras et al., 2007). in the last few years, h. tubulosa has been actively harvested in turkey, greece, italy, spain and the increasing of illegal and unregulated fishing is one of the main issues for its management (rakaj et al., 2019). overexploitation of this species led the italian ministry of agriculture, food and forestry (mipaaf) to ban sea cucumbers fishing along the entire national coastline (ministerial decree 156/2018), as a precaution for the conservation of the species. h. tubulosa is a continuous deposit-feeder, generally encountered in organic matter enriched soft bottoms and seagrass meadows (bulteel et al., 1992; gustato et al., 1982). coulon and jangoux (1993) reported that large individuals of h. tubulosa might ingest up to 17 kg of dry weight sediment ind-1 y-1. using the data provided by costa et al. (2014) it can be estimated that the quantity of seagrass detritus potentially ingested by h. tubulosa ranges between 12 and 28 g dry weight m−2 y−1 ind-1. the reproductive cycle of h. tubulosa was studied in specimens from the adriatic sea, oran coast (algeria) and dardanelles strait (turkey). the development stages of male and female gonads showed a clear annual pattern and all authors agreed that the spawning period was set between june and october with minor local differences, and a resting period from october to january (despalatovic et al., 2004; ocaña and tocino, 2005; dereli et al., 2015; tahri et al., 2019). rakaj et al. (2018) successfully bred and reared h. tubulosa in the laboratory, completing the larval development in 27 days, which, however, was followed by high mortality shortly after the settlement. a no nco mm er cia l u se on ly v. pasquini et al.40 recent study reported the use of h. tubulosa larvae as new model for embryo-larval bioassays to assess marine pollution (rakaj et al., 2021), but, to date, rearing techniques of this species remain still not very efficient. sea cucumbers in integrated multi-trophic aquaculture in the last two decades, to satisfy the demand for seafood product, aquaculture activities increased and the need to mitigate its impacts on the environment became an urgent need, especially in the presence of vulnerable habits like seagrass beds (pusceddu et al., 2007; holmer et al., 2008). wastes coming from mariculture plants can affect sediments biochemistry, increasing the organic contents, ultimately exacerbating eutrophication (david et al., 2009; keeley et al., 2014). in fact, wastes from mariculture can cause benthic hypoxia and anoxia, hydrogen sulphite enrichment and, in extreme cases, also led to rising of methanogenentic bacteria populations, which, in turn, can significantly impact the abundance and biodiversity of benthic organisms (karakassis et al., 2000; angel et al., 2002; mirto et al., 2002; burford et al., 2003; la rosa et al., 2004; fodelianakis et al., 2015). the conceptual approach of integrated multi-trophic aquaculture (imta) is to use different trophic-levels organisms in the same system: those belonging to the highest trophic level (generally fish) are fed artificially and those belonging to the lowest trophic level (extractive species) feed on waste released by the specimens of the highest trophic level (troell, 2009; granada et al., 2015). the extractive species commonly used in imta include molluscs, seaweeds or detritivorous species (zhou et al., 2006; slater and carton, 2007; yuan et al., 2013; slater et al., 2009; zamora and jeffs, 2011, 2012a, 2012 b; lamprianidou et al., 2015; shpigel et al., 2018). among detritivorous species, considering their feeding habits, sea cucumbers appear to be ideal candidates as extractive species for imta systems. commercially valuable holothurians species most used in imta systems include a. japonicus (zhou et al., 2006; yuan et al., 2013; kim et al., 2015), a. mollis (slater and carton, 2007; slater et al., 2009; zamora and jeffs, 2011, 2012a, 2012b), and p. californicus, (paltzat et al., 2008), mainly fed with scallops and mussels’ biodeposits alone, or mixed with powdered algae (yuan et al., 2006). other small-scale experiments used actinopyga bannwarthi panning, 1944 (israel et al., 2019) and h. scabra (mathieu-resuge et al., 2020). the imta feasibility in the mediterranean sea is still in an experimental scale, whereas either pilot or commercial scale activities have been carried out in other regions (macdonald et al., 2013; marinho et al., 2013; lamprianidou et al., 2015). to our best knowledge, only two studies investigated the use of h. tubulosa in imta systems in the mediterranean sea. beneath fish cages, tolon et al. (2017b) observed a biomass increase of holothurians ranging from 9 to 31 g ind-1 in just 90 days and suggested that these animals are ideal candidates to mitigate in imta the benthic eutrophication generated by fish farming. neofitou et al. (2019) during an experiment carried in the field beneath farming cages of the sea bream s. aurata and the sea bass dicentrarchus labrax linnaeus, 1758, reported that the maximum extractive capacity of holothurians is reached at a density of ca. 10 individuals m-2. such a density allowed abating om and oc contents in sediments beneath the cages by 31 and 59%, respectively. these results, though spatially and temporally fragmented, corroborate the idea of using sea cucumbers beneath fish cages, in imta systems, to mitigate the impacts of biodeposition on the sediment, at the same time providing a commercially important by-product, without any additional feed. with these assumptions, it can be envisaged that sea cucumbers in imta will increase the environmental sustainability of aquaculture and will also generate an important economic advantage, due to the high value of sea cucumbers. holothurians’ management perspectives the ecological consequences of holothurians overexploitation include a loss in bioturbation and a consequent reduction of benthic biomass, biodiversity, and ecosystem functioning (lohrer et al., 2004; solan et al., 2004; meysman et al., 2006b). therefore, sea cucumbers’ overexploitation claims for urgent measures to preserve natural populations and their ability to provide reproductive adults for either natural or artificial breeding. on the one hand, the peculiar biological and ecological traits of holothurians and the lack of reliable stock assessments make a scientific based management of this resource still far to be reached. management and regulation of sea cucumbers fishery are currently being implemented in some countries, using different approaches. among these, for example, a rotational zone strategy has been applied to the multispecies sea cucumber fishery in australia’s great barrier reef marine park, where this approach led to a substantial reduction of the risk of localized depletion, higher long-term yields, and improved economic performance (plagányi et al., 2015). to guarantee significant recruitment in an acceptable timeframe, future management policies of sea cucumbers should set a minimum population density threshold, below which exploitation should be banned (battaglene and bell, 2004), also to avoid the allee effect, which occurrence has been reported for overexploited populations of h. scabra in the warrior reef, australia (skewes et al., no nco mm er cia l u se on ly biology, ecology and management of deposit-feeders sea cucumbers 41 2000), i. fuscus in the galapagos marine reserve, ecuador (toral-granda and martinez, 2007), and h. nobilis (selenka 1867) in the great barrier reef, australia (uthicke and benzie, 2000) and holothuria mexicana ludwig, 1875 (rogers et al., 2018). ultimately, we notice that adequate protocols of holothurians’ populations management still need large amount of quantitative information about their population dynamics, recruitment success, rates of growth and natural mortality (romero-gallardo et al., 2018), mechanisms allowing larval settlement. concurrently, studies aiming at identifying new and more efficient ex situ rearing protocols, also to feed restocking actions and to preserve the natural genetic pools (purcell, 2004b, purcell and kirby, 2006; gonzález-wangüemert et al., 2015) are also needed. acknowledgments this study has been carried out in the framework of the projects: “marine habitats restoration in a climate changeimpaired mediterranean sea [mahres]”, funded by the ministero dell’istruzione dell’università e della ricerca under the prin 2017 call (protocol: 2017mhhwbn; cup f74i19001320001); “innovative species of commercial interest for sardinian aquaculture: development of experimental protocols for the breeding of sea cucumbers, (project n. 1/ina/2.47/2017)” founded by the european maritime and fisheries fund (emff) programme 2014/2020, measure: 2.47 – innovation; and co-founded by the “lneval: lncreasing echinoderm value chains” (grant n. id 101 ineval) founded by era-net bluebio programme. corresponding author: v.pasquini@studenti.unica.it key words: sea cucumber; holothuria tubulosa; biology; ecology; review. contributions: all the authors have read and approved the final version of the manuscript and agreed to be accountable for all aspects of the work. conflict of interest: the authors declare that they have no known competing financial interests or personal relationships that could have appeared to influence the work reported in this paper. availability of data and materials: all data generated or analyzed during this study are included in this published article. received: 21 july 2021. accepted: 12 octobre 2021. this work is licensed under a creative commons attribution noncommercial 4.0 license (cc by-nc 4.0). ©copyright: the author(s), 2021 licensee pagepress, italy advances in oceanography and limnology, 2021; 12:9995 doi: 10.4081/aiol.2021.9995 references agudo n, 2006. sandfish hatchery techniques. australian centre for international agricultural research, secretariat of the pacific community and worldfish center. noumea, new caledonia: 45 pp. amaro t, bianchelli s, billett dsm, cunha mr, pusceddu a, danovaro r, 2010. the trophic biology of the holothurian molpadia musculus: implications for organic matter cycling and ecosystem functioning in a deep submarine canyon. biogeosciences 7:2419-2432. anderson sc, flemming jm, watson r, lotze hk, 2011a. rapid global expansion of invertebrate fisheries: trends, drivers, and ecosystem effects. plos one 6:1-9. anderson sc, flemming jm, watson r, lotze hk, 2011b. serial exploitation of global sea cucumber fisheries. fish fish. 12:317-339. anderson sc, lotze hk, shackell nl, 2008. evaluating the knowledge base for expanding low trophic-level fisheries in atlantic canada. can. j. fish. aquat. sci. 65:2553-2571. andrew nl, agatsuma y, ballesteros e, bazhin ag, creaser ep, barnes dka, botsford lw, bradbury a, campbell a, dixon d, einarsson s, gerring pk, herbert k, hunter m, hur sb, johnson p k, juinio-meñez m a, kalvass p, miller rj, moreno ca, palleiro js, rivas d, robinson sm, schroeter sc, steneck rs, vadas rl, woodby da, xiaoqi z, 2002. status and management of world sea urchin fisheries. oceanogr. mar. biol. annu. rev. 40:343-425. angel dl, eden n, breitstein s, yurman a, katz t, spanier e, 2002. in situ bio-filtration: a means to limit the dispersal of effluents from marine finfish cage aquaculture. hydrobiologia 469:1-10. antoniadou c, vafidis d, 2011. population structure of the traditionally exploited holothurian holothuria tubulosa in the south aegean sea. cah. biol. mar. 52:171-175. battaglene sc, bell jd, 2004. the restocking of sea cucumbers in the pacific islands. in: bartley dm and leber, kl (eds.), case studies in marine ranching. fao fisheries technical paper 429:109-132. battaglene sc, seymour ej, ramofafia c, 1999. survival and growth of cultured juvenile sea cucumbers, holothuria scabra. aquaculture 178:293-322. berkes f, hughes tp, steneck rs, wilson ja, bellwood dr, crona b, folke c, gundersson lh, leslie hm, norberg j, nyström m, olsson p, österblom h, sheffer m, worm b, 2006. globalization, roving bandits, and marine resources. science 311:1557-1558. bordbar s, anwar f, saari n, 2011. high-value components and bioactives from sea cucumbers for functional food a review. mar. drugs. 9:1761-1805. bulteel p, jangoux m, coulon p, 1992. biometry, bathymetric distribution, and reproductive cycle of the holothuroid holothuria tubulosa (echinodermata) from mediterranean sea grass beds. mar. ecol. 13:53-62. burford ma, thompson pj, mcintosh rp, bauman rh, pearson dc, 2003. nutrient and microbial dynamics in high-intensity, zero-exchange shrimp ponds in belize. aquaculture 219:393-411. no nco mm er cia l u se on ly v. pasquini et al.42 çakly s, cadun a, kisla d, diçer t, 2004. determination of quality characteristics of holothuria tubulosa (gmelin 1788) in turkish sea (aegean region) depending on sun drying process step used in turkey. int. j. food sci. tech. 13:3-16. choo ps, 2008. population status, fisheries and trade of sea cucumbers in asia, p. 81-118. in: v. toral-granda, a. lovatelli and m. vasconcellos (eds.), sea cucumbers. a global review of fisheries and trade. no. 516. rome, italy. conand c, 2006. sea cucumber biology, taxonomy, distribution and conservation status, p.33-50. in a.w. bruckner (ed.), proceedings of the cites workshop on the conservation of sea cucumbers in the families holothuriidae and stichopodidae. noaa technical memorandum nmfsopr 34. silver spring, usa. conand c, polidoro b, mercier a, gamboa r, hamel jf, purcell s, 2014. the iucn red list assessment of aspidochirotid sea cucumbers and its implications. spc beche-de-mer inf. bull. 34:3-7. costa v, mazzola a, vizzini s, 2014. holothuria tubulosa gmelin 1791 (holothuroidea, echinodermata) enhances organic matter recycling in posidonia oceanica meadows. j. exp. mar. bio. ecol. 461:226-232. coulon p, jangoux m, 1993. feeding rate and sediment reworking by the holothuroid holothuria tubulosa (echinodermata) in a mediterranean seagrass bed off ischia island, italy. mar. ecol. prog. ser. 92:201-204. couvray s, miard t, bunet r, martin y, grillasca j-p, bonnefont j-l, coupé s, 2015. experimental release of paracentrotus lividus sea urchin juveniles in exploited sites along the french mediterranean coast. j. shellfish res. 34:1-9. dance sk, lane i, bell jd, 2003. variation in short-term survival of cultured sandfish (holothuria scabra) released in mangrove-seagrass and coral reef flat habitats in solomon islands. aquaculture 220:495-505. dar ma, ahmad ho, 2006. the feeding selectivity and ecological role of shallow water holothurians in the red sea. spc beche-de-mer inf. bull. 24:11-21. david cpc, maria yy, siringan fp, reotita jm, zamora pb, villanoy cl, sombrito ez, azanza rv, 2009. coastal pollution due to increasing nutrient flux in aquaculture sites. environ. geol. 58:447-454. dawbin wh. 1949. auto-evisceration and the regeneration of the viscera in the holothurian stichopus mollis (hutton). trans. r. soc. n.z. 77:497-523. dereli h, culha st, culha m, ozalp bh, tekinay aa, 2015. reproduction and population structure of the sea cucumber holothuria tubulosa in the dardanelles strait, turkey. med. mar. sci. 17: 47-55. despalatović m, grubelić i, šimunović a, antolić b, žuljević a, 2004. reproductive biology of the holothurian holothuria tubulosa (echinodermata) in the adriatic sea. j. mar. biol. assoc. uk. 84:409-414. dolmatov iy, 2014. asexual reproduction in holothurians. sci. world j. 2014:527234. dolmatov iy, 2021. molecular aspects of regeneration mechanisms in holothurians. genes (basel). 12:1-31. dolmatov iy, ginanova tt, 2009. post-autotomy regeneration of respiratory trees in the holothurian apostichopus japonicus (holothuroidea, aspidochirotida). cell. tissue. res. 336:41-58. domínguez-godino ja, gonzález-wangüemert m, 2018. breeding and larval development of holothuria mammata, a new target species for aquaculture. aquac. res. 49:1430-1440. domínguez-godino ja, slater mj, hannon c, gonzálezwangüermert m, 2015. a new species for sea cucumber ranching and aquaculture: breeding and rearing of holothuria arguinensis. aquaculture 438:122-128. duy ndq, francis ds, southgate pc, 2016. development of hyaline spheres in late auriculariae of sandfish, holothuria scabra: is it a reliable indicator of subsequent performance? aquaculture 465:144-151. eaves aa, richard palmer a, 2003. widespread cloning in echinoderm larvae. nature 425:146. fao, 2008. the state of world fisheries and aquaculture. technical report, food and agriculture organization of the united nations, rome, italy: 196 pp. fao, 2012. report on the fao workshop on sea cucumber fisheries: an ecosystem approach to management in the pacific (sceam pacific). food and agriculture organization of the united nations. report. no. 1003. rome, italy: 44 pp. fao, 2020. the state of world fisheries and aquaculture. sustainability in action. food and agriculture organization of the united nations. rome, italy: 224 pp. fodelianakis s, papageorgiou n, karakassis i, ladoukakis ed, 2015. community structure changes in sediment bacterial communities along an organic enrichment gradient associated with fish farming. ann. microbiol. 65:331-338. fredalina bd, ridzwan bh, abidin z, kaswandi m, zaiton h, zali i, kittakoop p, jais mm, 1999. fatty acid compositions in local sea cucumber, stichopus chloronotus, for wound healing. gen. pharmacol. 33:337-340. garcía-arrarás je, schenk c, rodrigues-ramirez r, torres ii, valentin g, candelaria ag, 2006. spherulocytes in the echinoderm holothuria glaberrima and their involvement in intestinal regeneration. dev. dyn. 235:3259-3267. giglioli aa, addis p, pasquini v, secci m, hannon c, 2021. first assessment of restocking efficacy of the depleted sea urchin paracentrotus lividus populations in two contrasted sites. aquac. res. 00:1-5. glockner-fagetti a, calderon-aguilera le, herrero-pérezrul md, 2016. density decrease in an exploited population of brown sea cucumber isostichopus fuscus in a biosphere reserve from the baja california peninsula, mexico. ocean coast. manag. 121:49-59. gonzález-wangüemert m, aydin m, conand c, 2014. assessment of sea cucumber populations from the aegean sea (turkey): first insights to sustainable management of new fisheries. ocean coast. manag. 92:87-94. gonzález-wangüemert m, domínguez-godino ja, cánovas f, 2018. the fast development of sea cucumber fisheries in the mediterranean and ne atlantic waters: from a new marine resource to its over-exploitation. ocean coast. manag. 151:165-177. gonzález-wangüemert m, valente s, aydin m, 2015. effects of fishery protection on biometry and genetic structure of two target sea cucumber species from the mediterranean sea. hydrobiologia 743:65-74. granada l, sousa n, lopes s, lemos mfl, 2015. is integrated multitrophic aquaculture the solution to the sectors’ major challenges? a review. rev. aquacult. 8:283-300. no nco mm er cia l u se on ly biology, ecology and management of deposit-feeders sea cucumbers 43 gustato g, villari a, del gaudio s, pedata p, 1982. [ulteriori dati sulla distribuzione di holothuria tubulosa, holothuria polii e holothuria stellati nel golfo di napoli.].[article in italian]. boll. soc. nat. napoli 91:1-14. hamel jjf, mercier a, 2008. population status, fisheries and trade of sea cucumbers in temperate areas of the northern hemisphere, p.257-291. in v. toral-granda, a. lovatelli and m. vasconcellos (eds.), sea cucumbers. a global review of fisheries and trade. fao fisheries and aquaculture technical paper. no. 516. rome, italy. hammond ls, 1983. nutrition of deposit-feeding holothuroids and echinoids (echinodermata) from a shallow reef lagoon, discovery bay, jamaica. mar. ecol. prog. ser. 10:297-305. hartati r, zainuri m, ambariyanto a, widianingsih w, 2020. feeding selectivity of holothuria atra in different microhabitat in panjang island, jepara (java, indonesia). biodiversitas 21:2233-2239. herrero-perezrul d, reyes-bonilla h, garcia-dominguez f, cintra-buenrostro ce, 1999. reproduction and growth of isostichopus fuscus in the southern gulf of california, mexico. mar. biol. 135:521-532. holmer m, argyrou m, dalsgaard t, danovaro r, diaz-almela e, duarte cm, frederiksen m, grau a, karakassis i, marba n, mirto s, perez m 2008. effects of fish farm waste on posidonia oceanica meadows: synthesis and provision of monitoring and management tools. mar. poll. bull. 56:1618-1629. horton t, kroh a, ahyong s, bailly n, boyko cb, brandão sn, et al. (2018) world register of marine species (worms). israel d, lupatsch i, angel dl, 2019. testing the digestibility of seabream wastes in three candidates for integrated multitrophic aquaculture: grey mullet, sea urchin and sea cucumber. aquaculture 510:364-370. işgören-emiroǧlu d, günay d, 2007. the effect of sea cucumber holothuria tubulosa (g., 1788) on nutrient and sediment of aegean sea shores. pak j biol sci. 10:586-589. ito s, kitamura h, 1997. induction of larval metamorphosis in the sea cucumber stichopus japonicus by periphitic diatoms. hydrobiologica 358:281-284. jumars pa, 2000. animal guts as non-ideal chemical reactors: partial mixing and axial variation in absorption kinetics. am. nat. 155:544-555. karakassis i, 2000. impact of cage farming of fish on the seabed in three mediterranean coastal areas. ices j. mar. sci. 57:1462-1471. keeley nb, macleod ck, hopkins ga, forrest bm, 2014. spatial and temporal dynamics in macrobenthos during recovery from salmon farm induced organic enrichment: when is recovery complete? mar. pollut. bull. 80: 250-262. kim t, yoon hs, shin s, oh mh, kwon i, lee j, choi sd, jeong ks, 2015. physical and biological evaluation of co-culture cage systems for grow-out of juvenile abalone, haliotis discus hannai, with juvenile sea cucumber, apostichopus japonicus (selenka), with cfd analysis and indoor seawater tanks. aquaculture 447:86-101. kinch j, purcell s, uthicke s, friederich k, 2008. population status, fisheries and trade of sea cucumbers in the western central pacific, p.7-55. in: toral-granda, m.v., lovatelli, a., vasconcellos, m. (eds.), sea cucumbers: a global review of fisheries and trade, fao fisheries and aquaculture technical paper. no. 516. rome, italy. koukouras a, sinis ai, bobori d, kazantzidis s, kitsos ms, 2007. the echinoderm (deuterostomia) fauna of the aegean sea, and comparison with those of the neighbouring seas. j. biol. res. 7:67-92. lamprianidou f, telfer t, ross lg, 2015. a model for optimization of the productivity and bioremediation efficiency of marine integrated multitrophic aquaculture. estuar. coast. shelf sci. 164:253-264. la rosa t, mirto s, mazzola a, danovaro r, 2001. differential responses of benthic microbes and meiofauna to fish-farm disturbance in coastal sediments. env. poll. 112:427-434. lee s, ford am, mangubhai s, wild c, ferse sca, 2018. effects of sandfish (holothuria scabra) removal on shallowwater sediments in fiji. peerj. 6:e4773. leiva g, castilla j, 2002. a review of the world marine gastropod fishery: evolution of catches, management and the chilean experience. rev. fish. biol. fish. 11:283-300. levinton, js, 1972. stability and trophic structure in depositfeeding and suspension-feeding communities. am. nat. 106:472-486. lohrer am, thrush sf, gibbs mm, 2004. bioturbators enhance ecosystem function through complex biogeochemical interactions. nature 431:1092-1095. macdonald cle, stead sm, slater mj, 2013. consumption and remediation of european seabass (dicentrarchus labrax) waste by the sea cucumber holothuria forskali. aquac. int. 21:1279-1290. mactavish t, stenton-dozey j, vopel k, savage c, 2012. deposit-feeding sea cucumbers enhance mineralization and nutrient cycling in organically enriched coastal sediments. plos one 7:50031. mangion p, taddei d, conand c, frouin p, 2004. feeding rate and impact of sediment reworking by two deposit feeders holothuria leucospilota and holothuria atra on fringing reef (reunion island, indian ocean), p.311-317. in: t. heinzeller and j. h. nebelsick (eds.), echinoderms: münchen, taylor and francis, london, uk. manini e, luna g, 2003. influence of the mineralogic composition on microbial activities in marine sediments: an experimental approach. chem. ecol. 19:399-410. mathieu-resuge m, le grand f, schaal g, kraffe e, lorrain a, letourneur y, lemonnier h, benoît, j, hochard s, 2020. assimilation of shrimp farm sediment by holothuria scabra: a coupled fatty acid and stable isotope approach. aquat. living resour. 33:1-12. mercier a, battaglene sc, hamel jf, 2000. settlement preferences and early migration of the tropical sea cucumber holothuria scabra. j. exp. mar. bio. ecol. 249:89-110. mercier a, hamel jf, 2009. endogenous and exogenous control of gametogenesis and spawning in echinoderms. adv. mar. biol. 55:1-302. meysman fjr, galaktionov os, gribsholt b, middelburg jj, 2006a. bioirrigation in permeable sediments: advective pore-water transport induced by burrow ventilation. limnol. oceanogr. 51:142-156. meysman fjr, middelburg jj, heip chr, 2006b. bioturbation: a fresh look at darwin’s last idea. trend. ecol. evol. 21:688–695. no nco mm er cia l u se on ly v. pasquini et al.44 mezali k, thandar as, 2014. first record of holothuria (roweothuria) arguinensis (echinodermata: holothuroidea: aspidochirotida: holothuriidae) from the algerian coastal waters. mar. biol. 7:40. michio k, kengo k, yasunori k, hitoshi m, takayuki y, hideaki y, hiroshi s. 2003. effects of deposit feeder stichopus japonicus on algal bloom and organic matter contents of bottom sediments of the enclosed sea. mar. pollut. bull. 47:118-125. mirto s, la rosa t, gambi c, danovaro r. mazzola a, 2002. nematode community response to fish-farm impact in the wester mediterranean. environ. pollut. 116:203-214. mohsen m, yang h, 2021. chapter 4 sea cucumbers research in the mediterranean and the red seas, p.61-101. in: m. mohsen, h. yang (eds.), sea cucumbers aquaculture, biology and ecology, academic press. moriarty djw, 1982. feeding of holothuria atra and stichopus chloronotus on bacteria, organic carbon and organic nitrogen in sediments of the great barrier reef. mar. freshw. res. 33:255-263. murray g, garcía-arrarás je, 2004. myogenesis during holothurian intestinal regeneration. cell. tissue res. 318:515-524. navarro, pg, garcía-sanz s, barrio jm, tuya f, 2013. feeding and movement patterns of the sea cucumber holothuria sanctori. mar. biol. 160:2957-2966. neofitou n, lolas a, ballios i, skordas k, tziantziou l, vafidis d, 2019. contribution of sea cucumber holothuria tubulosa on organic load reduction from fish farming operation. aquaculture 501:97-103. ocaña a, sanchez tocino, l, 2005. spawning of holothuria tubulosa (holothurioidea, echinodermata) in the alboran sea (mediterranean sea). zoologica baetica 16:147-150. pauly d, christensen v, guénette s, pitcher tj, sumaila u r, walters cj, watson r, and zeller d, 2002. towards sustainability in world fisheries. nature. 418:689-695. penry dl, jumars pa, 1986. chemical reactor analysis and optimal digestion. bioscience 36:310-315. penry dl, jumars, pa, 1987. modelling animals’ guts as chemical reactors. am. nat. 129:69-96 peters-didier j, sewell ma, 2019. the role of the hyaline spheres in sea cucumber metamorphosis: lipid storage via transport cells in the blastocoel. evodevo. 10:1-12. pietrak m, kim j, redmond s, kim y, yarish c, bricknell i, 2014. culture of sea cucumber in korea: a guide to korean methods and the local sea cucumber in the northeast u.s. maine sea grant college program. orono, usa: 12 pp. plagányi ée, skewes t, murphy n, pascual r, fischer m, 2015. crop rotations in the sea: increasing returns and reducing risk of collapse in sea cucumber fisheries. proc. natl. acad. sci. u. s. a. 112:6760-6765. purcell s, 2004a. criteria for release strategies and evaluating the restocking of sea cucumbers. in: lovatelli, a., conand, c., purcell, s.w., uthicke, s., hamel, j.-f., mercier, a. (eds.), advances in sea cucumber aquaculture and management. fao, dalian, china: 439 pp. purcell s, 2004b. rapid growth and bioturbation activity of the sea cucumber holothuria scabra in earthen ponds. proc. australasian aqua. 1:244. purcell s, simutoga m, 2008. spatio-temporal and size-dependent variation in the success of releasing cultured sea cucumbers in the wild. rev. fis. sci. 16:204-214. purcell sw, 2014. value, market preferences and trade of bechede-mer from pacific island sea cucumbers. plos one. 9:95075. purcell sw, conand c, uthicke s, byrne m, 2016. ecological roles of exploited sea cucumbers. oceanogr. mar. biol. annu. rev. 54:367–386. purcell sw, hair ca, mills dj 2012. sea cucumber culture, farming and sea ranching in the tropics: progress, problems and opportunities. aquaculture 368-369:68-81. purcell sw, kirby ds, 2006. restocking the sea cucumber holothuria scabra: sizing no-take zones through individualbased movement modelling. fish. res. 80:53-61. purcell sw, mercier a, conand s, hamel jf, toralgranda mv, lovatelly a, uthicke s, 2011. sea cucumber fisheries: global analysis of stocks, management measures and drivers of overfishing. fish fish. 14:34-59. purwati p, dwiono sap, 2005. fission inducement in indonesian holothurians. spc beche-de-mer inf. bull. 22:11-13. purwati p, dwiono, sap, 2007. experiment on fission stimulation of holothuria atra (holothuroidea, echinodemata): changing in body weight and morphology. mar. res. indonesia. 32:1-6. pusceddu a, fraschetti s, mirto s, holmer m, danovaro r, 2007. effects of intensive mariculture on sediment biochemistry. ecol. appl. 17:1366-1378. rakaj a, fianchini a, boncagni p, lovatelli a, scardi m, cataudella s, 2018. spawning and rearing of holothuria tubulosa: a new candidate for aquaculture in the mediterranean region. aquacult. res. 49:557-568. rakaj a, fianchini a, boncagni p, scardi m, cataudella s, 2019. artificial reproduction of holothuria polii: a new candidate for aquaculture. aquaculture 498:444-453. rakaj a, morroni l, grosso l, fianchini a, pensa d, 2021. towards sea cucumbers as a new model in embryo-larval bioassays: holothuria tubulosa as test species for the assessment of marine pollution. sci. total environ. 787:147593. ram r, chand rv, southgate pc, 2014. effects of processing methods on the value of bêche-de-mer from the fiji islands. j. mar. sci. res. dev. 4:1-7. ramírez-gonzález j, moity n, andrade-vera s, mackliff hr, 2020. estimation of age and growth and mortality parameters of the sea cucumber isostichopus fuscus (ludwig, 1875) and implications for the management of its fishery in the galapagos marine reserve. aquac. fish. 5:245-252. ramón m, simarro g, galimany e, lleonart j, 2019. evaluation of sediment particle size selection during feeding by the holothurian parastichopus regalis (cuvier, 1817). reg. stud. mar. sci. 31:100763. reise k 2002. sediment mediated species interactions in coastal waters. j. sea res. 48:127-141. reyes-bonilla h, herrero-pérezrul md, 2003. population parameters of an exploited population of isostichopus fuscus (holothuroidea) in the southern gulf of california, mexico. fish. res. 59:423-430. roberts d, gebruk av, levin v, manship bad, 2000. feeding and digestive strategies in deposit-feeding holothurians. oceanogr. mar. biol. ann. rev. 38:257-310. roberts d, moore hm, berges j, patching jw, carton mw, eardly df, 2001. sediment distribution, hydrolytic enzyme profiles and bacterial activities in the guts of oneirophanta no nco mm er cia l u se on ly biology, ecology and management of deposit-feeders sea cucumbers 45 mutabilis, psychropotes longicauda and pseudostichopus villosus: what do they tell us about digestive strategies of abyssal holothurians? prog. oceanogr. 50:443-458. romero-gallardo s, velázquez-abunader, lópez-rocha ja, garza-gisholt e, 2018. natural mortality estimates throughout the life history of the sea cucumber isostichopus badionotus (holothuroidea: aspidochirotida). peerj. 6:5235. schneider k, silverman j, kravitz b, rivlin t, schneider-mor a, barbosa s, byrne m, caldeira k, 2013. inorganic carbon turnover caused by digestion of carbonate sands and metabolic activity of holothurians. estuar. coast. shelf sci. 133:217-223. shiell g, 2004. field observations of juvenile sea cucumbers. spc beche-de-mer inf. bull. 20:11. shpigel m, shauli l, odintsov v, ben-ezra d, neori a, guttman l, 2018. the sea urchin, paracentrotus lividus, in an integrated multi-trophic aquaculture (imta) system with fish (sparus aurata) and seaweed (ulva lactuca): nitrogen partitioning and proportional configuration. aquaculture 490:260-269. sicuro b, levine j, 2011. sea cucumber in the mediterranean: a potential species for aquaculture in the mediterranean. rev. fish. sci. 19:299-304. skewes t, dennis d, burridge cm, 2000. survey of holothuria scabra (sandfish) on warrior reef, torres strait. cisro division of marine research. queensland, australia: 28 pp. slater mj, carton ag, 2007. survivorship and growth of the sea cucumber australostichopus (stichopus) mollis (hutton 1872) in polyculture trials with green-lipped mussel farms. aquaculture 207:389-398. slater mj, carton ag, 2009. effect of sea cucumber (australostichopus mollis) grazing on coastal sediments impacted by mussel farm deposition. mar. pollut. bull. 58:1123-1129. slater mj, jeffs ag, carton ag, 2009. the use of the waste from green-lipped mussels as a food source for juvenile sea cucumber, australostichopus mollis. aquaculture 292:219-224. slater mj, jeffs ag, sewell ma, 2011. selective movement and deposit-feeding in juvenile sea cucumber, australostichopus mollis determined in situ and in the laboratory. j. exp. mar. bio. ecol. 409:315-323. smiley s, mceuen fs, chaffee c, krishnan s, 1991. echinodermata: holothuroidea, p.663-750. in: giese, a.c., pearse, j.s., pearse, v.b. (eds.), reproduction of marine invertebrates, echinoderms and iophophorates, vol. vi. boxwood press, pacific grove, california, usa. solan m, cardinale bj, downing al, engelhardt, kam, ruesink jl, srivastava ds, 2004. extinction and ecosystem function in the marine benthos. science. 306:1177-1180. spirina is, dolmatov iy, 2003. mitotic activity in tissues of the regenerating respiratory tree of the holothurian apostichopus japonicus (holothuroidea, aspidochirota). russ. j. mar. biol. 29:123-125. strathmann rr, 1975. larval feeding in echinoderms. integr. comp. biol. 15:717-730. sun j, zhang l, pana y, lin c, wang f, kand r, yang h. 2015. feeding behavior and digestive physiology in sea cucumber apostichopus japonicus. physiol. behav. 139:336-343. tahri y, dermeche s, chahrour f, bouderbala m, 2019. the reproduction cycle of the sea cucumber holothuria (holothuria) tubulosa gmelin, 1791 (echinodermata holothuroidea holothuriidae) in oran coast, algeria. biodivers. j. 10:159-172. tolon mt, emiroglu d, gunay d, ozgul a, 2017a. sea cucumber (holothuria tubulosa gmelin, 1790) culture under marine fish net cages for potential use in integrated multitrophic aquaculture (imta). indian j. geo. mar. sci. 46:749-756. tolon t, emiroglu d, gunay d, hancı b, 2017b. effect of stocking density on growth performance of juvenile sea cucumber holothuria tubulosa (gmelin, 1788). aquac. res. 48:4124-4131. toral-granda mv, martínez pc, 2007. reproductive biology and population structure of the sea cucumber isostichopus fuscus (ludwig, 1875) (holothuroidea) in caamaño, galapagos islands, ecuador. mar. biodivers. 151:2091-2098. toral-granda v, 2008. galapagos islands: a hotspot of sea cucumber fisheries in central and south america, p.231-253. in v. toral-granda, a. lovatelli and m. vasconcellos (eds.), sea cucumbers. a global review of fisheries and trade. fao fisheries and aquaculture technical paper. no. 516. rome, fao. tortonese, e, 1965. fauna d’italia, echinodermata. ed. calderini, bologna, italia: 419 pp. troell m, 2009. chapter 9. integrated marine and brackishwater aquaculture in tropical regions: research, implementation and prospects, p.47-131. in d. soto (ed.), integrated mariculture: a global review. fao fisheries and aquaculture technical paper. no. 529. rome, fao. uthicke s, 2001. nutrient regeneration by abundant coral reef holothurians. j. exp. mar. bio. ecol. 265:153-170. uthicke s., benzie j.a.h. 2000. the effect of bêche-de-mer fishing on densities and size structure of holothuria nobilis (echinodermata: holothuroidea) populations on the great barrier reef. coral reefs. 19:271-276. uthicke s, benzie j, ballment e, 1998. genetic structure of fissiparous populations of holothuria (halodeima) atra on the great barrier reef. mar. biol. 132:141-151. uthicke s, benzie j, ballment e, 1999. population genetics of the fissiparous holothurian stichopus chloronotus (aspidochirotida) on the great barrier reef, australia. coral reefs. 18:123-132. uthicke s, conand c, 2005. amplified fragment length polymorphism (aflp) analysis indicates the importance of both asexual and sexual reproduction in the fissiparous holothurian stichopus chloronotus (aspidochirotida) in the indian and pacific ocean. coral reefs 24:103-111. uthicke s, conand c, benzie j, 2001. population genetics of the fissiparous holothurians stichopus chloronotus and holothuria atra (aspidochirotida): a comparison between the torres strait and la réunion. mar. biol. 139:257-265. uthicke s, karez r, 1999. sediment patch selectivity in tropical sea cucumbers (holothuroidea: aspidochirotida) analysed with multiple choice experiments. j. exp. mar. biol. ecol. 236:69-87. uthicke s, purcell s, 2004. preservation of genetic diversity in restocking of the sea cucumber holothuria scabra planned through allozyme electrophoresis. can. j. fish. aquat. sci. 61:519-528. wen j, hu c, fan s 2010. chemical composition and nutritional quality of sea cucumbers. j. sci. food agric. 90:2469-2474. no nco mm er cia l u se on ly v. pasquini et al.46 wilkie ic, 2001. autotomy as a prelude to regeneration in echinoderms. microsc. res. tech. 55: 369-396. yanagisawa t, 1998. chapter 9. aspects of the biology and culture of the sea cucumber, p.291-308. in: s. de silva (ed.), tropical mariculture. academic press, london, uk. yang h, bai y, 2015. chapter 1. apostichopus japonicus in the life of chinese people, p.1-24. in: h. yang , j.f. hamel, a. mercier (eds.), the sea cucumber apostichopus japonicus. history, biology, and aquaculture. academic press. yang h, hamel jf, mercier a, 2015. the sea cucumber apostichopus japonicus: history, biology and aquaculture. elsevier, london, uk: 454 pp. yuan x, yang h, meng l, wang l, li y, 2013. impacts of temperature on the scavenging efficiency by the deposit-feeding holothurian apostichopus japonicus on a simulated organic pollutant in the bivalve-macroalage polyculture from the perspective of nutrient budgets. aquaculture 406-407:97-104. yuan x, yang h, zhou y, mao y, zhang t, liu y, 2006. the influence of diets containing dried bivalve feces and/or powdered algae on growth and energy distribution in sea cucumber apostichopus japonicus (selenka) (echinodermata: holothuroidea). aquaculture 256:457-467. zamora ln, jeffs ag, 2011. feeding, selection, digestion and absorption of the organic matter from mussel waste by juveniles of the deposit-feeding sea cucumber, australostichopus mollis. aquaculture 317:223-228. zamora ln, jeffs ag, 2012a. feeding, metabolism and growth in response to temperature in juveniles of the australasian sea cucumber, australostichopus mollis. aquaculture 358359: 92-97. zamora ln, jeffs ag, 2012b. the ability of the deposit-feeding sea cucumber australostichopus mollis to use natural variation in the biodeposits beneath mussel farms. aquaculture 326-329:116-122. zang y, tian x, dong s, dong y, 2012. growth, metabolism and immune responses to evisceration and the regenerationof viscera in sea cucumber, apostichopus japonicus. aquaculture 358-359:50-60. zhang x, sun l, yuan j, sun y, gao y, zhang l, li s, dai h, hamel jf, liu c, yu y, 2017. the sea cucumber genome provides insights into morphological evolution and visceral regeneration. plos biol. 15:2003790. zhou y, yang h, liu s, yuan x, mao y, zhang t, liu y, zhang f, 2006. feeding on biodeposits of bivalves by the sea cucumber stichopus japonicus selenka (echinodermata: holothuroidea) and a suspension co-culture of filter-feeding bivalves with deposit feeders in lantern nets from longlines. aquaculture 256:510-520. no nco mm er cia l u se on ly