115 

Journal of Multidisciplinary Applied Natural Science Vol. 2 No. 2 (2022) 

Research Article 

Removal of Toxic Cations from Aqueous Solutions using 
Ginger Root Waste 

Jude Chinedu Onwuka*, Stephen Igberi Azubuike, and Timothy M. Akpomie 

Received : April 4, 2022  Revised : May 26, 2022  Accepted : May 28, 2022 Online : May 29, 2022 

Abstract
Recently, the harmful impact of toxic metals in the aquatic environment cannot be over emphasized again. This work investigated 
the potential application of ginger root waste (GRW) to remove toxic cations (Cd

2+ 
and Pb

2+
) from the aqueous medium. Batch 

adsorption examination was carried out as a function of sorbent dose, initial metal ion concentration, contact time, and temperature. 
The sorption equilibrium of the metal ions onto the GRW was subjected to Langmuir, Freundlich, Elovich and Redlich-Peterson 
isotherm models over concentration ranges of 10-50 mg/L. Sorption information was used for kinetic and thermodynamic 
modeling. The GRW materials before and after sorption was characterized using FTIR and SEM. Results showed higher removal 
percentage of Cd

2+
 over Pb

2+
 ions in all the factors studied. The Redlich – Peterson isotherm model affirmed that sorption of Cd

2+
 

and Pb
2+

 occurred in a heterogenous surface of the sorbent which is strongly influenced by multiple micropores and caves. Kinetic 
studies revealed that the sorption was controlled through intra-particle diffusion model aided by surface and chemical reactions. 
Meanwhile, thermodynamic parameters indicated that the Cd

2+
 and Pb

2+
 sorption process was endothermic, however, non-

spontaneous at temperature of 303 and 313 K. The FTIR and SEM data showed the evidence of successful sorption of the toxic 
cations on to the sorbent material.  
 
Keywords Ginger root waste, sorption, toxic cations, kinetics, thermodynamics

1. INTRODUCTION

Heavy metal contamination is a serious 

challenge to the human race and the natural 

ecosystem. The environmental hazard and their 

effects on human existence are very serious [1] due 

to their accumulation in the human body and 

inability to biodegrade into harmless waste [2]. 

Today, the dangerous global impact of these metals 

on the ecosystem has become a non-push-over 

reality that is very difficult to be circumvented. 

Environmental pollution by heavy metals occurs by 

means of industrialization and extraction of natural 

resources [3]. Most industrial discharge and effluent 

contain high amounts of heavy metals which 

consequently constitute serious concern due to the 

high toxicity it portends to the living organism [4]. 

Organic wastes are quite susceptible to 

biodegradation, but heavy metals are not simply 

degraded by any biological means into harmless by-

 
products [2].  

The global increase in industrialization has 

accelerated industrial applications of heavy metals 

mainly as the catalyst and thus, increased their 

presence in wastewater giving rise to environmental 

pollution and wastewater contamination. Among 

the heavy metals, cadmium and lead, which are 

quite carcinogenic, are often found in industrial 

effluents such as those related to mining, 

metallurgy, petrochemicals, electroplating, 

batteries, pharmaceutical industries, paper, pulp, 

plastic, and paint manufacturing [5]. These metals 

are presumed toxic since they are not metabolized 

by the human body rather, they keep accumulating 

in the human tissues [6]-[9]. Thus, it becomes 

incumbent to search for the possible way of 

removing these toxic metals from the aqueous 

medium. 

As a result of the many health challenges 

associated with contamination of water bodies by 

heavy metals, a lot of techniques such as chemical 

precipitation, ion exchange, electro-dialysis, 

membrane separation, and nanofiltration have been 

implored in preventing the presence of metal 

cations in our drinking water samples. However, 

there are obvious limitations inherent with the 

application of these techniques [10] including less 

efficiency and sensitivity, generation of sludge, 

high maintenance cost, high energy and reagent 

requirement, and inability to sequestrate metal ions 

at low concentration [11]. These limitations 

Copyright Holder: 

© Onwuka, J. C., Azubuike, S. I., and Akpomie, T. M. (2022) 
 

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J. Multidiscip. Appl. Nat. Sci. 

116 

Table 1. Data isolated from equilibrium adsorption isotherm plots. 

Toxic 
cation 

           Langmuir Isotherm        Freundlich Isotherm          Elovich Isotherm        Redlich-Peterson 

qmax b RL R
2
 Kf n R

2
 qm Ke R

2
 β A R

2
 

Cd
2+

 -0.9832 -0.1709 0.8562 0.7691 0.0463 0.3493 0.9640
*
 -3.2927 1.5095 0.8868 1.8622 0.0463 0.9187 

Pb
2+

 4.0016 0.0646 0.7946 0.0045 11.1970 -1.0092 0.0495 1.6292 5.4016 0.9817
*
 1.9906 11.1931 0.1735 

Note:
 
* means highest R

2
 values. R

2
 values that are moderate are bolded 

prompted individuals and research institutes to 

engage themselves in finding the best viable 

technique to adequately bring this menace into 

limits acceptable by standard organizations and one 

of such options is well provided in the adsorption 

process [12]. The adsorption process is cheap, eco-

friendly and cost-effective as this process does not 

produce sludge and does not require high energy 

and reagents [10][11]. 

Globally, safe drinking water is everyone's 

concern, as it is difficult to find anybody who has 

not been affected by waterborne ailment [13]. 

Termed as a “universal solvent”, a higher 

percentage of the global population depends on 

water for drinking and domestic uses [10]. Due to 

the indispensable daily need and use of water across 

the globe, it has been made vulnerable and 

susceptible to being contaminated by the heavy 

metal discharge of domestic and industrial 

effluents, radioactive waste, mining waste, leakages 

from oil pipelines and atmospheric deposition [14]. 

Consequently, this has made water one of the key 

sources of global infection and diseases [15]. 

Statistics show that 80% of diseases and infections 

globally are waterborne and there is the propensity 

that it will be doubled if nothing is done in the near 

future [16]. At the moment, the global concern on 

heavy metal accumulation in drinking water and its 

effects on the human body is on the increase. The 

apprehension is that no biological means of 

degradation have been identified for these set of 

elements [3]. 

Health experts opined those natural products 

such as ginger drink are of great benefit to the 

human body [17]. Recent discoveries showed that it 

prevents nausea and vomiting, muscle pain and 

curbs cancer growth, soothes sore throat, and 

prevents common cold [18][19]. Nutritionally, the 

ginger drink is rich in antioxidants which are 

necessary for combating activities of free radicals in 

the body [12]. According to Adanlawao and Dairo 

[20], the ginger drink is quite rich in potassium, an 

important mineral for biochemical reactions in the 

body. Chronic intake of synthetic beverages 

majority of which are soft drinks is susceptible to 

dental erosion, obesity with a high risk of 

cardiovascular dysfunction and type 2-diabetes 

[18]. Due to these obvious health concerns on 

synthetic drink consumption, the focus has been 

shifted towards the consumption of natural drinks 

such as ginger root drinks [21]. Unfortunately, after 

the aqueous extraction of the ginger root, the wastes 

are dumped indiscriminately, thus, constituting an 

environmental nuisance. Previous research [12] 

utilized ginger root in removing selected toxic 

metal cations. However, this research aims to study 

the potential of using ginger root waste (GRW) to 

specifically remove Cd
2+

 and Pb
2+

 from the aqueous 

solution.     

 

2. MATERIALS AND METHODS 

 

2.1. Sample Collection and Preparation  

The fresh ginger roots used for this research 

(201 g) were obtained from the Lafia main market, 

in Nasarawa State. It was transported in a 

polyethylene bag to the Department of Chemistry 

Laboratory, Federal University of Lafia. It was 

washed and rinsed severally with deionized water to 

remove dust and adhering soil particles. The thin 

outer covers were carefully peeled using stainless 

steel knife. The sample was washed again and 

sliced into bits and pounded using a ceramic mortar. 

The pounded sample was placed in a stainless 

basin, and 2 L of deionized water was added. The 

mixture was allowed to stand for 10 min, stirred and 

then filtered. The residue obtained after the juice 

extraction was re-soaked in 2 L of deionized water 

for another 10 min. This process was repeated for 

further three consecutive times and filtered after 

each soaking. The chaff obtained after this series of 

extraction is referred to as the ginger root waste. 

The waste was air-dried for 24 h and then oven-

dried for 72 h at 80 °C. The dried sample was 



J. Multidiscip. Appl. Nat. Sci. 

117 

ground into fine powder using ceramic mortar and 

pestle and sieved to less than 1 mm fine particle 

size with a sieve and stored in a plastic sealable bag 

ready for the chemical analysis.  

 

2.2. Batch Sorption Studies 

The experimental solution was prepared by 

diluting the stock solution earlier prepared with 

deionized water. A 50 mL portion of the prepared 

10 mg/L metal ion solution was measured into 250 

mL plastic container and 0.5 g of the sorbent 

(GRW) was added. The plastic container was closed 

and then placed on a shaker (Bernareggio MI, PID 

system type M 428-Bd) with a temperature-

controlled water bath. It was shaken continuously 

for 60 min at 27 °C and 120 rpm. The supernatant 

was filtered using 0.2 mm Whatman filter paper. 

The filtrate was analyzed using atomic absorption 

spectrophotometry (PG990) in order to ascertain the 

equilibrium metal ion concentrations. Various 

parameters of the adsorption process were studied 

as a function of sorbent dose, initial metal ion 

concentration, contact time and temperature of the 

solution. The effect of sorbent dose (0.5, 1.0, 1.5, 

2.0, and 2.5 g) was studied by fixing the volume (50 

mL) and concentration (10 mg/L) of toxic cations as 

well as the contact time (60 min) and temperature 

(27 °C). The effect of initial metal ion concentration 

(10, 20, 30, 40, and 50 mg/L) was carried out by  

metal ions sorption at a constant sorbent dose, 

contact time and temperature. The contact time 

effect (30, 60, 90, 120, and 150 min) was studied 

under various temperatures (303, 313, 323, and 343 

K) while keeping other factors constant. Similarly, 

the variation of temperature (303, 313, 323, and 343 

K) was carried out at various contact times (30, 60, 

90, 120 and 150 min). The amount of metal ion 

sorbed per unit mass was calculated using a mass 

balance as shown Eq. 1. 

 

      (1) 

 

Where qe is heavy metal ion concentration 

absorbed onto the sorbent at equilibrium (mg/g), Ci 

and Ce are initial and equilibrium concentrations 

(mg/L) of the metal ions, respectively, m is mass 

(mg) of sorbent, and V is the volume (L) of sorbate 

solution.  

 

The adsorption or removal percentage (%) was 

calculated using Eq. 2. 

 

      (2) 

 

2.3. Adsorption Isotherm Modeling 

This experiment was used to establish the 

relationship between the amount of the toxic cations 

(Cd
2+

 and Pb
2+

) that was adsorbed by the GRW 

sorbent GRW and the equilibrium concentrations of 

the toxic cations at a constant temperature [12]. 

Langmuir, Freundlich, Elovich and Redlich-

Peterson isotherm models were used to evaluate 

how the toxic cations (adsorbate) interacted with 

the GRW sorbent. This study was conducted by 

fitting the obtained data from the effect of initial 

metal ion concentration into the isotherm models. 

 

2.3.1. Langmuir Isotherm 

The equation was developed by Irving Langmuir 

(1916). The linearized form of the Langmuir model 

as shown in Eq. 3, was used in this study.  

 

Figure 1. Effect of sorbent dose on Cd
2+ 

and Pb
2+

 

sorption.  

Figure 2. Effect of initial metal ion concentration 

on Cd
2+ 

and Pb
2+

 sorption.  



J. Multidiscip. Appl. Nat. Sci. 

118 

 

      (3) 

 

Where Ce is concentration at equilibrium or final 

concentration of the metal in solution (mg/L), qe is 

adsorption efficiency or the amount of the adsorbate 

adsorbed at equilibrium (mg/g), qmax is the 

maximum adsorption capacity (mg/g), and b is the 

Langmuir constant related to the energy of 

adsorption (L/mg). A plot of the graph Ce/qe against 

Ce, produces 1/qmax as the slope, and thus, qmax was 

calculated. The value of b which is the Langmuir 

constant was evaluated from the intercept of the 

graph which is 1/b x qmax [22]. 

Dimensionless equilibrium function known as 

RL value was evaluated using Eq. 4 expressed as: 

 

      (4)  

 

When RL = 1, a linear adsorption occurred. If RL 

= 0, there is a irreversible adsorption. But when 0< 

RL < 1, then a favorable adsorption occurred. 

 

2.3.2. Freundlich Isotherm 

The linear form of Freundlich isotherm 

demonstrated in Eq. 5, was applied. 

 

      (5) 

 

A plot of logqe against logCe produced 1/n as the 

slope, from where n was evaluated, while from the 

intercept (logkf), kf was estimated. It should be 

noted that kf and n are Freundlich constants related 

to adsorption capacity and adsorption intensity, 

respectively [23]. 

 

Figure 3.  Effect of contact time for Cd
2+

 sorption 

at various temperatures. 

Figure 4.  Effect of contact time for Pb
2+

 sorption at 

various temperatures. 

2.3.3. Elovich Isotherm 

The linear form of Elovich isotherm used is 

given as Eq. 6. 

 

      (6) 

 

The plot of  against  gives a straight line. 

Thus, Elovich maximum adsorption capacity (qm) 

and Elovich constant (Ke) were calculated from the 

slope (   )and intercept (         ), respectively. 

 

2.3.4. Redlich-Peterson Isotherm 

The linear form of the Redlich-Peterson 

isotherm used is shown in Eq. 7.  

 

      (7) 

 

Where A is the Redlich-Peterson isotherm 

constant (L/g), β (L/mg) is the exponent, Ce (mg/L) 

is the liquid-phase equilibrium concentration of the 

sorbent (GRW) while qe is the sorbate (Cd
2+

 and 

Pb
2+

) equilibrium onto the GRW sorbent. When the 

graph of InCe/qe vs InCe is plotted, the slope is β 

while the intercept is lnA. When β value is below or 

equal to 1, the Redlich-Peterson isotherm is 

confined to Langmuir isotherm, however, when β 

value is above 1, the Redlich-Peterson isotherm 

tends to Freundlich isotherm. The constant “A” 

indicates the sorption capacity [24][25]. 

 

2.4. Kinetic Studies  

This was studied to measure the uptake of the 

toxic cations as a function of time at a constant 

concentration so as to determine the mechanism or 

reaction pathway involved in the toxic cation 

uptake by the sorbent. Thus, data obtained from the 



J. Multidiscip. Appl. Nat. Sci. 

119 

effect of contact time were analyzed using pseudo–

first order, pseudo–second order, intra–particle 

diffusion and liquid film diffusion models. 

 

2.4.1. Pseudo First–Order kinetic 

Pseudo first-order kinetic was carried out to 

establish the contribution of physisorption or 

surface reaction, to the uptake of the toxic cations 

by the sorbent. According to Amadi et al. [26], the 

linear form of pseudo first–order kinetics is given 

below. 

 

      (8) 

 

The plot of log(qe-qt) against time ‘t’, gave - 

as the slope and logqe as the intercept [27]. 

 

2.4.2. Pseudo Second–Order Kinetic 

The contribution of chemisorption was evaluated 

by analyzing the data with pseudo-second order 

kinetic model. The linear form of the pseudo-second 

order kinetic is stated Eq. 9. 

 

      (9) 

 

The plot of against t, gives k2 as the slope 

and as the intercept. 

 

2.4.3. Intra–Particle Diffusion 

The equation for the linear form of intra-particle 

diffusion kinetics is given Eq 10. 

 

qt = kd × t
1/2

 + C               (10)   

 

Thus, when qt is plotted against t
1/2

, the slope is 

the rate constant kd in (mg/(g min
1/2

)) and the 

intercept ‘C’ in (mg/g). The importance of the 

intercept is that it explains the thickness of the 

boundary of the layer. The characteristic 

significance of the value of the constant ‘C’ is that 

the larger the value, the greater the boundary layer 

thickness [22][27]. 

 

2.4.4. Liquid–Film Diffusion 

This was used to investigate the rate at which the 

toxic cations in the form of a film glide over one 

another [25]. The linear form of liquid film 

diffusion model is given below. 

 

In(1-α) = klft + C             (11) 

 

Where α = qt/qe, klf is the liquid film diffusion 

rate constant. A plot of In(1-α) against time (t) 

yields the constant klf (min
-1

) as the slope and a 

dimensionless constant C as the intercept [27].  

 

2.5. Thermodynamic Studies 

Thermodynamic parameters such as enthalpy 

change (∆H), entropy change (∆S), and Gibbs’ free 

energy (∆G), of the adsorption processes were 

determined. The Arrhenius equation described by 

Onwuka et al. [27], was used for the 

thermodynamic modeling, the equation is expressed 

in Eq. 12. 

 

               (12) 

 

The enthalpy change (∆H) was determined from 

the slope of the plot of         versus       using Eq.12. 

The heat capacity (Cp) and entropy change, 

(∆S), were determined from Eqs. (13) and (14) 

respectively. 

 

               (13) 

 

               (14)  

 

The Gibbs’ free energy was determined using 

Eq. 15. 

 

   Source of Variation SS df MS Fcal P-value Fcrit 

Cd
2+

 Between Groups 10.46042 4 2.615106 0.005798 0.999927 2.866081 

  Within Groups 9021.483 20 451.0741    

  Total 9031.943 24     

Pb
2+

 Between Groups 2178.819 4 544.7049 1.127699 0.371616 2.866081 

  Within Groups 9660.465 20 483.0232       

  Total 11839.28 24         

Table 2. ANOVA for effect of contact time on the sorption of Cd
2+

 and Pb
2+

 at various temperatures. 



J. Multidiscip. Appl. Nat. Sci. 

120 

  Source of Variation SS df MS F P-value Fcrit 

Cd
2+

 Between Groups 175150.1 5 35030.03 1247.75 3.31705E-22 2.772853 

  Within Groups 505.3421 18 28.07456    

  Total 175655.5 23         

Pb
2+

 Between Groups 189251.5 5 37850.31 277.7577 2.22831E-16 2.772853 

  Within Groups 2452.877 18 136.2709       

  Total 191704.4 23         

Table 3. ANOVA for effect of temperature for Cd
2+

 and Pb
2+

 sorption at different time intervals. 

 

               (15) 

 

Where  is the amount adsorbed at temperature 

T, is the amount adsorbed at temperature   ,            

R is the universal gas constant,  is the heat 

capacity of the system at pressure p,     and  

are the initial and final pressures respectively, 

and  are the initial and final temperatures, 

respectively. 

 

2.6. Characterization 

 

2.6.1. Fourier Transform Infra-Red (FT-IR) 

Analysis 

The KBr pellet was prepared by mixing about 1 

mg of the sorbent with 250 mg KBr (FT-IR grade), 

pressed and placed for acquisition of IR spectrum 

on SHIMADZU model (8400S) of the FT-IR 

equipment. The spectra were recorded between 

4000–400 cm
-1

 resolutions of 4 cm
-1

. This was 

carried out at the National Research Institute for 

Chemical Technology (NARICT) Zaria, Kaduna 

State. 

 

2.6.2. Scanning Electron Microscope (SEM) 

Analysis 

The morphology of the GRW sorbent before and 

after metal ion sorption, was investigated using 

Scanning Electron Microscope (SEM; JSM-

5610LV). The micrographs were taken at 1000 

times magnification. 

 

3. RESULTS AND DISCUSSIONS 

 

3.1. Effect of Sorbent Dose 

The result of the effect of sorbent dose on the 

metal ions (Cd
2+

 and Pb
2+

) removal efficiency is 

presented in Fig. 1. The result shows that the 

sorption capacity is greatly influenced by the 

amount of sorbent material in the aqueous solution 

since the percentage sorption of the toxic metal ions 

increased in a higher sorbent dose. This suggests 

that there are more active binding sites, pores and 

caves at higher amount of the sorbent as similarly 

reported by Kampalanonwat and Supaphol [28] in a 

related investigation of heavy metal ion adsorption 

by animated polyacrylonitrile nanofibre mats. 

Increasing the sorbent dose also reduces the 

distance metal ions needed to travel in order to be 

entrapped by active pores and caves. A higher 

amount of the sorbent in solution results in a 

proportionate increase in the surface area which 

enhanced the sorption of more metal cations. A 

similar observation was reported by Vijayakurma et 

al. [29] and Shooto et al. [12] on natural pelite and 

ginger root, respectively. It was also observed in 

Fig 1 that increment in sorbent dose leads to higher 

Cd
2+

 adsorption through a maximum while Pb
2+

 

adsorption increased continuously with the range of 

sorbent dose studied. This suggests that Cd
2+

 

adsorption attained equilibrium faster than Pb
2+

 

adsorption within the studied sorbent dosages. This 

can be attributed to the smaller ionic size of Cd
2+

 

(compared to Pb
2+

) which makes it more mobile 

and consequently results in more interaction with 

the active sites on the sorbent. 

 

3.2. Effect of Initial Metal ion Concentrations 

Fig. 2 shows the effect of initial metal ions 

concentration on the sorption of the toxic metal ions 

from the aqueous medium. It is observed that the 

removal of the metal ions increased with increase in 

the initial metal ions concentration. This suggests 

that at lower metal ion concentration, desorption of 

these metal cations from the sorbent occurs more 

than their adsorption on the sorbent. It also 

indicates that the metal ions interacted more with 



J. Multidiscip. Appl. Nat. Sci. 

121 

Temp 

(K) 

Pseudo First–order Pseudo Second–order Intra–particle Diffusion Liquid Film Diffusion 
qexp 

K1 R
2
 qcal K2 R

2
 qcal Kd R

2
 C Klf R

2
 C 

Cd
2+

 

303 0.012 0.086 0.132 0.540 0.282 -0.458 0.006 0.583
a
 1.769 0.011 0.069 2.506 1.847 

313 0.009 0.072 0.145 0.264 0.381 0.278 0.006 0.540
a
 1.769 0.009 0.051 2.447 1.862 

323 0.007 0.053 0.129 0.301 0.178 0.205 0.005 0.383
a
 1.794 0.008 0.040 2.542 1.874 

333 0.007 0.050 0.132 0.291 0.157 0.204 0.005 0.340
a
 1.798 0.007 0.034 2.558 1.881 

Pb
2+

 

303 0.000 0.004 0.205 0.026 0.002 0.081 0.040 0.772
a
 0.954 0.000 0.000 2.002 1.456 

313 0.021 0.470 1.189 0.221 0.480 -0.135 0.065 0.633
a
 0.993 0.021 0.375 0.275 1.749 

323 0.014 0.173 0.308 0.296 0.291 -0.879 0.022 0.590
a
 1.548 0.015 0.139 1.627 1.810 

333 0.002 0.017 0.252 0.018 0.062 0.229 0.008 0.113 1.744 0.002 0.010 1.900 1.927 

Table 4. Data Isolated from Kinetic Plots. 

Note:
 a
 means highest R

2
 values. R

2
 values that are moderate are bolded  

the sorbent active binding sites and caves at their 

higher metal ion concentrations compared to lower 

metal ion concentrations [30]. This can also imply 

that the ionic orientation favours sorption at higher 

initial metal ions concentration which is ascribed to 

the fact that the force of repulsion at higher 

concentrations has a minimum barrier to the 

sorption of the metal cations [31].  

It was also observed in Fig. 2 just as in Fig. 1, 

that the sorption of Cd
2+

 from an aqueous medium 

was higher than that of Pb
2+

. This disparity can be 

attributed to the differences in the atomic/ionic 

radius between the two toxic cations. Egila et al. 

[32], in similar research on Amarathus hydridus L. 

stalk waste, reported that atoms with lower atomic 

radii show more tendencies to be sorbed easily than 

atoms with larger atomic radii. The atomic radii of 

Cd
2+

 is 0.97 Å while that of Pb
2+

 is 1.20 Å. Thus, 

this made it relatively difficult for Pb
2+

 ions to 

infiltrate the lignocellulosic matrix component of 

the GRW sorbent and hence, result in lower Pb
2+

 

sorption compared to Cd
2+

 sorption. Chukwuemeka 

et al. (2015), Onwuka et al. [27] and Shooto et al. 

[12] reported similar findings in their works on 

ginger, Delonix regia pods and ginger root, 

respectively.   

 

3.3. Adsorption Isotherm Studies 

The nature of the interaction between the toxic 

cations and the sorbent was assessed using 

Langmuir, Freundlich, Elovich and Redlich–

Peterson isotherm models. Parameters evaluated 

and isolated from the isotherm plots in Equations 3, 

5, 6 and 7 are summarized in Table 1. Generally, 

the value of coefficient of determination (R
2
) in 

Table 1 shows that the order of best fit isotherm 

model for Cd
2+

 sorption onto the GRW sorbent is 

Freundlich > Redlich–Peterson > Elovich > 

Langmuir isotherm while that of Pb
2+

 sorption is 

Elovich > Redlich–Peterson > Freundlich > 

Langmuir isotherm. 

RL value is the most important feature of the 

Langmuir adsorption isotherm that helps to explain 

and predict the degree of interaction and 

compatibility between the sorbate and sorbent in the 

aqueous solution [22][33]. Favourable Langmuir 

isotherm modeling is such that 0<RL<1 as observed 

by Kussel et al. [34]. Thus, it was observed (Table 

1) that the Langmuir dimensionless constant (RL) 

values for Cd
2+

 and Pb
2+

 sorption from the aqueous 

medium, is less than 1 suggesting that the sorption 

of these toxic cations
 
onto the GRW sorbent is 

relatively favourable. The high Langmuir R
2
 value 

for Cd
2+

 sorption further supports this assertion.  

However, the low Langmuir R
2
 value for Pb

2+
 

sorption and the negative qmax for Cd
2+

 sorption 

which influenced the outcome of few other 

variables shown in Table 1, are indications that the 

application of Langmuir isotherm model alone 

cannot satisfactorily explain the sorption of Pb
2+

 

and Cd
2+

 onto the sorbent [32][35][36]. 

In the Freundlich isotherm studies, the “n” value 

explains the degree of compatibility and interaction 

between the metal ions and the sorbent, while Kf is 

the adsorption capacity. The best fit “n” value for 

Freundlich adsorption isotherm is that value that 



J. Multidiscip. Appl. Nat. Sci. 

122 

falls between 0 and 1 [37]. On this basis, it can be 

inferred that the sorption of Cd
2+

 onto the sorbent 

favourably aligned with the Freundlich isotherm 

modeling. This entails that the sorption is multilayer 

in nature and a physical process. This finding 

agrees with reports by Jimoh et al. [38], Amadi et 

al. [26] and Lei et al. [39] on Cd
2+

 adsorption using 

modified charcoal, cellulosic biosorbent derived 

from okra stem and modified biomass ash, 

respectively. The high R
2
 value (Table 1) 

strengthens the assertion that Cd
2+

 sorption onto the 

GRW is favourable to the Freundlich isotherm. 

However, the negative “n” value of the Freundlich 

isotherm model for Pb
2+

 sorption implies that its 

sorption was quite difficult and the application of 

Freundlich isotherm model alone is inadequate to 

satisfactorily explain the sorption process. The 

difficulty in Pb
2+

 sorption could be attributed to its 

large ionic radius or the charge density [32]. This 

observation was opined by Alsenani [35] in similar 

research on Calligonum comosum leaf powder.  

Table 1 further illustrates that Cd
2+

 and Pb
2+

 

sorption onto the sorbent, aligned with the Elovich 

isotherm model based on their high coefficient of 

determination (R
2
) values. This suggests that the 

sorption sites increased exponentially with 

adsorption validating that sorption of these toxic 

cations (Cd
2+

 and Pb
2+

) is multilayer in nature [40]. 

It is therefore inferred that the sorption of Cd
2+

 and 

Pb
2+

 was dependent on the multiple micropores and 

the complicated chemical structure of sorbent. This 

was also reported by Achmad et al. [41] and Lei et 

al. [39] in similar works on a novel adsorbent 

developed from Uncaria gambir extract and 

adsorption of Cd
2+

 from aqueous solution by 

modified biomass ash. 

Redlich–Peterson isotherm is used to correct the 

errors encountered in using Langmuir and 

Freundlich adsorption isotherm [42]. The model is 

actually the blend of Langmuir and Freundlich 

adsorption isotherms into one [24][43]. The Redlich

–Peterson isotherm modeling for the Cd
2+

 sorption 

was also presented in Table 1. The high coefficient 

of determination (R
2
) shows that Redlich-Peterson 

isotherm model satisfactorily predicted the sorption 

of Cd
2+

 onto the sorbent. It was also observed from 

Table 1 that the values of β for Cd
2+

 (1.8622) and 

Pb
2+

 (1.9906) sorption are above unity. This implies 

minimal or little influence of Langmuir adsorption 

isotherm on Cd
2+

 and Pb
2+

 sorption onto the 

sorbent.  This assertion is anchored on the findings 

of Belhachemi and Addoun [24] that adsorption 

process where β approaches unity is suggestive that 

the sorption is best fitted to Langmuir isotherm 

model but if it is above unity, the sorption is best 

explained by the Freundlich isotherm model. Thus, 

it is inferred that the sorption of Cd
2+

 and Pb
2+

 

satisfactorily aligned to Freundlich isotherm model. 

Figure 5. Effect of temperature for Cd
2+

 sorption at 

different contact time.   

Figure 6. Effect of temperature for Pb
2+

 sorption at 

different contact time. 

Metal 

ion 

Cp 

(J/K) 

∆H 

(J/mol) 

∆S 

(J/K) 

∆G 

(J/mol) 

T 

(K) 

Cd
2+

 32.092 320.092 1.011  14.587 303 

           4.477 313 

          -5.633 323 

        -15.743 333 

Pb
2+

 1267.636 12,676.365 39.931 577.260 303 

        177.950 313 

        -221.360 323 

        -620.667 333 

Table 5. Thermodynamic data of Cd
2+

 and Pb
2+

 

sorption.  



J. Multidiscip. Appl. Nat. Sci. 

123 

The implication is that the sorption of Cd
2+

 and Pb
2+

 

occurred on a heterogeneous surface and is strongly 

influenced by multiple micropores, caves and rough 

surface morphology of the sorbent which constitutes 

the active sorption sites. This observation is 

validated by the strong agreement between the 

sorption capacities for Redlich–Peterson isotherm 

(A = 0.0463 for Cd
2+

 and 11.1931 for Pb
2+

) and 

Freundlich adsorption isotherm (Kf = 0.0463 for 

Cd
2+

 and 11.1970 for Pb
2+

) as shown in Table 1. 

Gimbert et al. [44] reported similar findings on 

catonized starch-based material in a single 

component system. 

 

Figure 7. FTIR spectra of (a) unloaded (b) Cd
2+

 loaded and (c) Pb
2+ 

loaded sorbents. 



J. Multidiscip. Appl. Nat. Sci. 

124 

3.4. Effect of Contact Time at Various 

Temperatures 

The effect of contact time on Cd
2+

 and Pb
2+

 

sorption are shown Figs 3 and 4. The results show 

that the sorption of toxic metal ions (Cd
2+

 and Pb
2+

) 

at various temperatures studied, increased through a 

maximum with increase in contact time (Figs 3 and 

4). Thus, the sorption of toxic metal ions was very 

fast initially at the adsorbent-adsorbate contact until 

equilibrium (saturation) was attained at 90 min and 

then it gradually started decreasing despite the 

increase in the contact time. This is attributed to the 

saturation of the available pores, caves and 

adsorptive active binding sites. This observation 

was also reported by Vijayakumar et al. [29] in 

similar work on natural pelite. 

It was further observed that a two-way 

adsorption process played out. First, was the rapid 

phase encounter that occurred between the time of 

the initial adsorbate-adsorbent contact to the time 

equilibrium was reached and the second phase was 

the slower phase that started from the time 

equilibrium was reached downward with declining 

adsorption even with increasing contact time. This 

is ascribed to the fact that at saturation point, all 

active sites, pores and caves are filled up and a 

further increase in the contact time results in 

desorption. This accounts for the decline in the 

sorption process after equilibrium has been reached, 

an ascertion also shared by Kekpugile et al. [10] in 

a related research on kinetic study of garlic and 

ginger. A similar inference was also opined by 

Kampalanonwat and Supaphol [28] in a related 

work on animated polyacrylonitrile nanofibre mats. 

Also, higher percentage removal of Cd
2+

 over Pb
2+

 

was also observed (Figs 3 and 4). This is an 

indication that Cd
2+ 

is more mobile and dynamic 

than Pb
2+

 in solution which is a function of the 

differences in the ionic size and charge density. 

Jimoh et al. [38] reported similar findings in related 

work on modified charcoal.  

 

3.5. One–way factorial Analyses 

Figs 3 and 4 showed the variations in the 

removal efficiency of Cd
2+

 and Pb
2+

 with contact 

time at different temperatures respectively. These 

variations in the removal efficiency of the toxic 

metal cations with time were tested for significant 

differences. The analysis of variance (ANOVA) 

presented in Table 2 shows that the observed 

variations in the sorption of  Cd
2+

 and Pb
2+

 with 

contact time have no significant difference at the 

various temperatures studied. This is because the p-

values are greater than α = 0.05 and Fcal is less than 

Fcrit.  Thus, it is a further validation of the null 

hypothesis. This result implies that any of the 

studied temperature can be used to explain/describe 

the variations of the metal ions sorption with 

contact time. 

In Figs 5 and 6, the variations in the sorption of 

Cd
2+

 and Pb
2+

 with temperature at various contact 

time were shown. Table 3 shows that there is 

statistically significant difference in Cd
2+

 and Pb
2+

 

sorption when the temperature was varied at 

different time. It can be concluded that at α = 0.05, 

null hypotheses were rejected because Fcal is higher 

than Fcrit and p-values is less than α = 0.05. This is 

an indication that thermal influence is essential for 

the sorption of these metal adsorbates as was also 

observed by Batool et al. [45] in related study on 

linear and non-linear approach and error analysis. 

 

3.6. Kinetic Studies  

Reaction (pseudo first–order and pseudo second

–order) and diffusion (Intra–particle and liquid film 

diffusion) kinetics were used to determine the 

sorption mechanism of the toxic cations (Cd
2+

 and 

Pb
2+

) on the sorbent. The kinetic data extracted 

from the plots in Equations 7, 8, 9 and 10, are 

summarized in Table 4. The values of coefficient of 

determination (R
2
), qcal and qexp for the various 

kinetic plots, were used to determine the 

mechanisms that contributed to the sorption of the 

toxic cations at each temperature. It was observed 

that pseudo first–order and pseudo second–order 

kinetic models could not satisfactorily explain the 

sorption process of the metal cations at all the 

temperatures studied (Table 4). This is because 

there is no strong agreement between the qcal and 

qexp values [22].  

The values of coefficient of determination (R
2
) 

within 0.43 ≤ R
2 

≤ 0.83 are considered high and 

moderate [46]. The highest values of R
2
 are 

obtained at intra–particle diffusion plots of all 

studied temperatures. Hence, diffusion into the 

pores of the GRW sorbent is the rate-controlling 

mechanism in the sorption of the toxic cations (Cd
2+

 

and Pb
2+

) at all the studied temperatures (303–333 



J. Multidiscip. Appl. Nat. Sci. 

125 

K). However, intra–particle diffusion is not the sole 

mechanism in the sorption of these toxic cations 

because the linear plot did not pass through the 

origin [10][22]. It was also observed that chemical 

reaction contributed to the sorption mechanism of 

Cd
2+

 and Pb
2+

 at 313 K while surface reaction and 

transport of sorbate molecules from liquid to solid 

phase were majorly involved in Pb
2+

 sorption 

mechanism at 313 K. The intercepts obtained in the 

intra–particle diffusion plots are relatively high in 

all the sorption temperatures for Cd
2+

 sorption and 

at 323 K and 333 K for Pb
2+

 sorption. This indicates 

that the sorption process of these metal ions was 

mainly a surface reaction [26][27]. Other related 

studies that shared similar views are those of 

Ademiluyi and Ujile [47], Li et al. [48] and 

Kulkarni et al. [49].  

 

3.7. Effect of Temperature at Different Time 

Intervals 

The effect of temperature on the sorption Cd
2+

 

and Pb
2+

 at different contact times by the sorbent is 

shown in Figs 5 and 6 respectively. The result 

indicated that the sorption of the toxic cations (Cd
2+

 

and Pb
2+

) increased steadily as the temperature was 

raised. This is attributed to the activated mobility of 

the metal cations introduced into the system by 

increase in temperature and also, the diffusion of 

the ionic species [38]. This is believed in 

facilitating the sorption of these toxic cations as 

demonstrated by the increase in percentage removal 

of the metal cations in the aqueous solution [50]. 

Therefore, thermal influence is opined to have a 

strong effect on the sorption of the metal ions. This 

was similarly reported by other kinetic studies by 

Wase [51] and Tounsadi et al. [36] for tea factory 

waste and Diplotaxis hara glebionis corona L. 

 

3.8. Thermodynamic Studies 

The data isolated from the thermodynamic plots 

using Equation 11, are summarized in Table 5. In 

the result, it was observed that the heat capacity of 

Cd
2+

 sorption onto the sorbent is 32.09 J/mol while 

that of Pb
2+

 sorption was 1267.636 J/mol. Heat 

capacity Cp is the quantity of thermal energy 

required to raise the temperature of adsorption by 

one degree [52]. The higher heat capacity of 

adsorption of Pb
2+

 can be attributed to the higher 

charge density of Pb
2+

 compared to Cd
2+

. This is 

believed to have made it less mobile and so 

required higher thermal activation for the reaction 

to occur [38][53].  

Table 5 further showed that the sorption of the 

metal cation was non-spontaneous at low 

temperatures of 303–313 K as depicted by the 

positive value of change in Gibbs free energy (∆G). 

However, as the sorption temperature increased to 

323–333 K, the values of ∆G for the metal cation 

sorption become negative suggesting that the meta 

cation sorption is spontaneous. The implication is 

that for sorption of these metal cations to occur onto 

the sorbent, heat must be added. This is ascertained 

by the positive value of enthalpy change (∆H) 

which indicates that the sorption of these metal 

cations is endothermic in nature. Therefore, for the 

reaction (metal cation sorption) proceed, the initial 

energy barrier is required to be broken by increase 

in temperature. This is contrary to the findings of 

Figure 8. SEM Micrographs of (a) unloaded (b) Cd
2+

 loaded and (c) Pb
2+ 

loaded sorbents. 



J. Multidiscip. Appl. Nat. Sci. 

126 

Shooto et al. [12] but in agreement with findings of 

Kazi et al. [54], Hussein et al. [52] and Gabreslassie 

[55]. 

 

3.9. Fourier Transform Infra-red (FT-IR) 

Spectroscopy  

The FTIR analysis of the unloaded, Cd
2+

 loaded 

and Pb
2+

 loaded sorbents, are presented in Fig 7a, 

7b and 7c, respectively. The peak 2337.80 cm
-1

 

appeared in both the unloaded and metal ions 

loaded sorbents. This peak falls within 2400-2000 

cm
-1

 and thus,  assigned to O=C=O stretching. The 

peak 1643.41 cm
-1

 as seen in Cd
2+

 and Pb
2+

 loaded 

sorbents (Fig 7b and 7c) falls within 1648-1638 cm
-

1
 and is assigned to C=C stretching from alkene. 

The peak 1735.99 cm
-1

 falls within 1735-1750 cm
-1

 

is assigned to C=O stretching.  As the sorbent was 

hydrolysed during the metal ions sorption, the peak 

3356.25 cm
-1

 appeared in both Fig 7b and 7c. This 

peak falls within the region 4000–3000 cm
-1

 which 

is assigned to O–H (stretching) resulting from 

ROH. This peak is very important in the 

investigation by virtue of the nature of sorbent used 

for the analysis. The GRW material is a plant fibre 

[56] and commonly fibre is a soluble polymer in 

solution containing hydrophilic groups such as the 

O–H groups that undergoes such a reaction 

mechanism that can entrap divalent cations [57]. 

Figs 7a, 7b and 7c show that during sorption, the 

bivalent toxic cations (Cd
2+

 and Pb
2+

) interacted 

with the O–H, C=C and C=O groups which caused 

spectra enhancement/shift and also spectra 

appearance. The interaction between these 

functional groups and the bivalent toxic ions led to 

the sorption of these bivalent toxic cations [58][12].  

 

3.10. Scanning Electron Microscope (SEM) 

Analysis 

Figs 8a, 8b and 8c show the SEM micrographs 

of the unloaded, Cd
2+

 loaded and Pb
2+

 loaded 

sorbents, respectively. Pores, caves and roughened 

surfaces were observed in Fig 8a. These are factors 

suspected to facilitate and influence the sorption of 

Cd
2+

 and Pb
2+

 onto the sorbent [45][59]. Figs 8b 

and 8c show that after the sorption of the metal 

ions, there was conglutination of the neighboring 

fibres at the touching line making the morphology 

of the sorbent compact and the sorbent appeared 

smoother and slippery with tiny blocks depicting 

the distribution of the metal ions on the sorbent 

surface. This is evidence that sorption of these 

metal ions (Cd
2+

 and Pb
2+

) onto the sorbent was 

successful which is in agreement with the previous 

studies [28][60]. 

 

4. CONCLUSIONS 

 

This study showed that the ginger root waste 

(GRW) adsorbent material has the potential to 

remove toxic cations (Cd
2+

 and Pb
2+

) from the 

aqueous medium. The GRW material was able to 

remove more Cd
2+

 from aqueous medium than Pb
2+

. 

The sorption of these toxic cations occurred on a 

heterogeneous surface and is strongly influenced by 

multiple micropores, caves and rough surface 

morphology of the sorbent. Diffusion into the 

sorbent pores was pivotal for the sorption of these 

toxic cations. The sorption process is non–

spontaneous at low temperature but spontaneous at 

higher temperature. The FTIR results showed that 

these toxic cations (Cd
2+

 and Pb
2+

) were adsorbed at 

both hydrophilic and hydrophobic functional 

groups. SEM micrographs showed that the 

adsorption of these toxic cations changed the 

surface morphology of the GRW sorbent. 

 

AUTHOR INFORMATION 

 

Corresponding Author 

Jude Chinedu Onwuka — Department of 

Science Laboratory Technology, Federal 

University of Lafia, Lafia – 950101 (Nigeria); 

Department of Chemistry, Federal University of 

Lafia, Lafia – 950101 (Nigeria); 

orcid.org/0000-0003-2950-8273 

Email: emperor20062003@yahoo.com 

 

Authors 

Stephen Igberi Azubuike — Department of 

Chemistry, Federal University of Lafia, Lafia – 

950101 (Nigeria); 

Timothy M. Akpomie — Department of 

Chemistry, Federal University of Lafia, Lafia – 

950101 (Nigeria); 

 

ACKNOWLEDGEMENT 

The authors are immensely grateful to Mr. 

William Egbe and other staff and management of 

https://orcid.org/0000-0003-2950-8273
mailto:emperor20062003@yahoo.com


J. Multidiscip. Appl. Nat. Sci. 

127 

National Soil Testing Laboratory Complex, Federal 

Ministry of Agriculture and Rural Development, 

Goningora, Kaduna State; for availing their 

laboratory services for this research. 

 

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