# Removal of Zn (II) and Cu (II) Ions from Industrial Wastewaters Using Magnetic Biochar Derived from Water Hyacinth

Authors: Benias C. Nyamunda; Terrence Chivhanga; Upenyu Guyo; Fidelis Chigondo
Journal: Journal of Engineering
 (2019)

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## Abstract

This study evaluates the effectiveness of magnetic biochar (Fe2O3-EC) derived from water hyacinth in the removal of Cu+2 and Zn+2 from aqueous solution. Fe2O3-EC was prepared by chemical coprecipitation of a mixture of FeCl2 and FeCl3 on water hyacinth biomass followed by pyrolysis. The adsorbent was characterized by Fourier transform infrared spectroscopy (FTIR) and scanning electron microscopy (SEM) coupled with energy dispersive spectroscopy (EDX). Batch adsorption studies on the effects of temperature, biosorbent dosage, contact time, and initial metal ion concentration were carried out. Fe2O3-EC exhibited optimum contact time, biosorbent dosage, and pH values of 65 min, 1.2 g, and 6, respectively. Fe2O3-EC exhibited strong magnetic separation ability and high sorption capability. Metal ion adsorption onto the biochar conformed to the Langmuir isotherm. Kinetic studies revealed that the adsorption process followed pseudo-second-order model. The calculated thermodynamic parameters showed that the adsorption process was feasible and exothermic in nature. These results have demonstrated that the use of Fe2O3-EC in metal ion removal could provide an alternative way to manage and utilize this highly problematic invasive species.


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## Body

## 1. Introduction

Rapid increase in human industrial activities such as agriculture, mining, and manufacturing has resulted in discharge of heavy metal contaminated wastewaters [1, 2] into main water bodies. The nonbiodegradability of toxic heavy metals results in their bioaccumulation in aquatic plants and animals [3, 4]. Among toxic heavy metal pollutants, zinc and copper are often present in industrial wastewaters. These heavy metals originate from various industrial processes such as electroplating, textile, photographic materials, painting, galvanizing, battery, paper, and explosives manufacturing [5]. There is need to develop efficient and cost effective ways of removing heavy metal contaminants from industrial wastewaters.Several industrial wastewater techniques such as ion exchange, solvent extraction, chemical precipitation, and activated carbon adsorption are usually applied to remove heavy metal pollutants [6–8]. However these technologies suffer drawbacks of incomplete pollutant removal, high reagent or energy requirements, and toxic sludge generation [9]. Biosorption has been reported as an emerging, effective, competitive, and inexpensive technology because of its high efficiency and easy handling nature for heavy metal ion removal [10]. Adsorption of heavy metals on surface of biomaterials occurs through different functional groups such as amino, carboxyl, hydroxyl, phenolic, ester, sulfhydryl, and phosphate [9]. Several plant based biosorbents have been used for the removal of heavy metal pollutants from wastewaters [10–13]. These include eucalyptus bark [14], bush mango [15], lignocellulose [16], marula [17], rice husk [18], and switchgrass [19]. Huge volumes of polluted industrial wastewaters still demand exploration of newer adsorbents.Water hyacinth (Eichhornia Crassipes) is ranked amongst the most prevalent aquatic weeds exhibiting global adverse effects on aquatic ecosystems [20, 21]. The renowned impact of uncontrolled growth of water hyacinth includes disruption of water abstraction facilities and blockage of irrigation networks [22, 23]. The weed is also a favourable breeding ground for health hazard vectors such as mosquitos and bilharzia snails.To overcome problems associated with water hyacinth, several weed management methods such as physical removal, chemical application of herbicides, and use of biological control agents have been adopted [24]. However these methods are expensive, ineffective, and labour and energy intensive [25–27].Although water hyacinth is responsible for numerous challenges, it has found wide ranging useful applications as a phytoremediation agent of industrial wastewaters [28]. One effective way for improved management of this highly invasive species involves conversion of water hyacinth into biochar for the treatment of industrial effluents.Biochar is a carbon-rich product derived from pyrolysis of biomass with little oxygen. Compared to activated carbon, biochar pyrolysed at temperatures not more than 700°C has been demonstrated to possess higher functional groups and mineral ash [29] rendering it a potential environmental sorbent for various ionic pollutants [30–32]. Biochar is a promising low cost adsorbent for metal ion removal from industrial wastewaters. Its application however is limited by low metal sorption capacities [33]. To enhance metal sorption capacities, biochars have been modified by loading with minerals, organic functional groups, reductants, nanoparticles, and alkali solution [34].The use of biochar derived from hyacinth in the removal of phosphates [35], arsenate [36], cadmium [37], and chromium [38] has been reported in literature. Coprecipitation of iron salts on water hyacinth followed by pyrolysis improved potential of the biochar for efficient removal of these ionic pollutants. However to the best of our knowledge no study has been reported for the use of hyacinth biochar in the removal of Cu+2 and Zn+2 from wastewaters.This study evaluates the effectiveness of biochar derived from water hyacinth in removing Cu+2 and Zn+2 from aqueous solution. An investigation of the kinetics and isotherm studies of these metal ions for the biochar sorption will be done.


## 2. Method

### 2.1. Preparation of Biosorbent

Water hyacinth was collected from Lake Chivero located about 22 km south of Harare, Zimbabwe. The water hyacinth sample was chopped followed by drying in sunlight for 5 days. The samples were further dried in an oven set at 100°C for 24 h. The dried material was ground and sieved to produce particles in size range 0.25–1.75 mm. Ground water hyacinth samples were kept in an air tight container. The dried biomass (200 g) was saturated in 1000 mL aqueous equimolar (2 M) mixture of FeCl2 and FeCl3. Under vigorous magnetic stirring, 5 M NaOH solution was added dropwise to increase the suspension pH to 10. The mixture was subjected to continuous stirring for 30 min prior to separation by centrifugation at 3000 rpm for 10 min. A ceramic pot was tightly filled with the biomass. The full pot was covered with a lid prior to pyrolysis. During pyrolysis, the biomass was initially heated to 100°C for 2 h. This was followed by heating at a rate of 5°C per min to attain three maximum temperatures (200, 300, and 600°C) and continuing heating for 6 h. The pyrolysed products were cooled to room temperature followed by grinding and sieving to produce uniform particle sizes (250 μm). The biochar particles were rinsed with deionized distilled water several times to remove most organic matter. The saturated biomass was then dried in an oven.


### 2.2. Characterisation of Biochar

Physisorption measurements of Fe2O3-EC and EC were carried out using Micromeritics ASAP 2020 surface area and porosity analyzer. Samples (0.05 g) were degassed to eliminate any carbon dioxide and water, so that a clean surface was obtained. The samples were then submitted to physisorption analysis for 12 h. Analyses of the superficial element composition of Fe2O3-EC and imaging of the surface morphology were done on a scanning electron microscope (SEM) using a JEOL-7500 field emission electron microscope at 20 keV. Attached to it was an energy dispersive X-ray spectrometer (EDX). The samples were mounted onto metal stubs with double sided carbon tape before imaging. The FTIR of the adsorbent were obtained using Thermofisher Scientific spectrometer (Nicolet 6700, USA) in the frequency range 500-4000 cm−1. Magnetisation curve of biochars was obtained using PPMS 9T vibrating sample magnetometer (Quantum design, USA) at room temperature under applied field range -15000 to 15000 Oe.


### 2.3. Batch Adsorption Experiments

Stock solutions (1000 mg L−1) of Cu+2 and Zn+2 were prepared by dissolving appropriate amounts of nitrate precursor salts (Sigma, South Africa) in deionized water and acidified with concentrated HNO3 (5 mL). Biosorption experiments were carried out in tightly closed 250 mL Erlenmeyer flasks containing 100 mL of separate metal ion solutions. The flasks containing 1 g biochar were continuously agitated (150 rpm) at room temperature on a shaker for 2 h. The resultant solutions were filtered on Whatman 42 filter paper, and the supernatant liquid was analyzed on a flame atomic absorption spectrometer (FAAS), Shimadzu AA-680, Japan. The percentage of metal ion removal was calculated using (1)R=Co-Ce100Co.The biosorption efficiency was calculated according to the following equation:(2)qe=Co-CeVMwhere qe is the amount (mg g−1) of metal ion adsorbed by biochar, Co and Ce are the metal ion concentrations (mg L−1) in the solution initially and after biosorption, respectively, V is the volume (L) of the solution, and m is the mass (g) of biosorbent. The influence of pH on the removal of metal ions was monitored in the pH range 2-10. pH adjustments were done using aqueous NaOH or HCl. The effects of other parameters on metal ion adsorption were monitored at different parameter ranges: temperature (25-50°C), biosorbent dosage (0.2-2 g), contact time (30-125 min), and initial metal ion concentration (20-100 mg L−1).


### 2.4. Desorption and Reusability Studies

Biosorbent samples (1.2 g) of Fe2O3-EC and EC were contacted with 100 mL aqueous solutions of 15 mg L−1 Cu+2 or Zn+2 at 25°C and pH 6 for 65 min. After metal ion sorption, the adsorbents were filtered and washed thrice with deionized water to remove residual metal ions on surface. The adsorbents were then kept in contact with 100 mL NaOH solutions (0.1-0.5 M) to desorb metal ions. The mixtures were shaken on a rotary shaker for 15 min. The filtrates were analyzed for metal ion content on FAAS.


## 2.1. Preparation of Biosorbent

Water hyacinth was collected from Lake Chivero located about 22 km south of Harare, Zimbabwe. The water hyacinth sample was chopped followed by drying in sunlight for 5 days. The samples were further dried in an oven set at 100°C for 24 h. The dried material was ground and sieved to produce particles in size range 0.25–1.75 mm. Ground water hyacinth samples were kept in an air tight container. The dried biomass (200 g) was saturated in 1000 mL aqueous equimolar (2 M) mixture of FeCl2 and FeCl3. Under vigorous magnetic stirring, 5 M NaOH solution was added dropwise to increase the suspension pH to 10. The mixture was subjected to continuous stirring for 30 min prior to separation by centrifugation at 3000 rpm for 10 min. A ceramic pot was tightly filled with the biomass. The full pot was covered with a lid prior to pyrolysis. During pyrolysis, the biomass was initially heated to 100°C for 2 h. This was followed by heating at a rate of 5°C per min to attain three maximum temperatures (200, 300, and 600°C) and continuing heating for 6 h. The pyrolysed products were cooled to room temperature followed by grinding and sieving to produce uniform particle sizes (250 μm). The biochar particles were rinsed with deionized distilled water several times to remove most organic matter. The saturated biomass was then dried in an oven.


## 2.2. Characterisation of Biochar

Physisorption measurements of Fe2O3-EC and EC were carried out using Micromeritics ASAP 2020 surface area and porosity analyzer. Samples (0.05 g) were degassed to eliminate any carbon dioxide and water, so that a clean surface was obtained. The samples were then submitted to physisorption analysis for 12 h. Analyses of the superficial element composition of Fe2O3-EC and imaging of the surface morphology were done on a scanning electron microscope (SEM) using a JEOL-7500 field emission electron microscope at 20 keV. Attached to it was an energy dispersive X-ray spectrometer (EDX). The samples were mounted onto metal stubs with double sided carbon tape before imaging. The FTIR of the adsorbent were obtained using Thermofisher Scientific spectrometer (Nicolet 6700, USA) in the frequency range 500-4000 cm−1. Magnetisation curve of biochars was obtained using PPMS 9T vibrating sample magnetometer (Quantum design, USA) at room temperature under applied field range -15000 to 15000 Oe.


## 2.3. Batch Adsorption Experiments

Stock solutions (1000 mg L−1) of Cu+2 and Zn+2 were prepared by dissolving appropriate amounts of nitrate precursor salts (Sigma, South Africa) in deionized water and acidified with concentrated HNO3 (5 mL). Biosorption experiments were carried out in tightly closed 250 mL Erlenmeyer flasks containing 100 mL of separate metal ion solutions. The flasks containing 1 g biochar were continuously agitated (150 rpm) at room temperature on a shaker for 2 h. The resultant solutions were filtered on Whatman 42 filter paper, and the supernatant liquid was analyzed on a flame atomic absorption spectrometer (FAAS), Shimadzu AA-680, Japan. The percentage of metal ion removal was calculated using (1)R=Co-Ce100Co.The biosorption efficiency was calculated according to the following equation:(2)qe=Co-CeVMwhere qe is the amount (mg g−1) of metal ion adsorbed by biochar, Co and Ce are the metal ion concentrations (mg L−1) in the solution initially and after biosorption, respectively, V is the volume (L) of the solution, and m is the mass (g) of biosorbent. The influence of pH on the removal of metal ions was monitored in the pH range 2-10. pH adjustments were done using aqueous NaOH or HCl. The effects of other parameters on metal ion adsorption were monitored at different parameter ranges: temperature (25-50°C), biosorbent dosage (0.2-2 g), contact time (30-125 min), and initial metal ion concentration (20-100 mg L−1).


## 2.4. Desorption and Reusability Studies

Biosorbent samples (1.2 g) of Fe2O3-EC and EC were contacted with 100 mL aqueous solutions of 15 mg L−1 Cu+2 or Zn+2 at 25°C and pH 6 for 65 min. After metal ion sorption, the adsorbents were filtered and washed thrice with deionized water to remove residual metal ions on surface. The adsorbents were then kept in contact with 100 mL NaOH solutions (0.1-0.5 M) to desorb metal ions. The mixtures were shaken on a rotary shaker for 15 min. The filtrates were analyzed for metal ion content on FAAS.


## 3. Results

### 3.1. Characterisation of the Adsorbent

Table1 shows the specific surface areas of Fe2O3-EC and EC at different pyrolysis temperatures. Surface areas of EC were smaller than those of Fe2O3-EC. This could be attributed to Fe2O3-EC containing considerable proportion of iron oxide particles, which have small surface areas and abundant transitional pores [39]. For both adsorbents, an increase in surface area with increasing temperature was observed due to volatilization of transitional fluxes (secondary reactions) presented by the thermal decomposition of the biomass cellulosic and semicellulosic structures [40]. These secondary reactions become prominent with increasing temperature resulting in continued cracking of the biomass structure forming voids on the surface. However, as temperature was further increased beyond 500°C, there was a notable decrease in surface area due to collapse of the overall mesoporous structure of the adsorbents as a result of volatilization of most of the biomass inorganic binding sites [40].Table 1
Specific surface areas of EC and Fe2O3-EC.
Temperature (°C) Adsorbent Surface area (m2 g−1) 350 EC 150.13 Fe2O3-EC 163.60  450 EC 610.88 Fe2O3-EC 1038.48  600 EC 553.59 Fe2O3-EC 983.75The formation of iron oxides could be explained by the fact that during pyrolysis the ferric chloride underwent drastic dehydration process forming Fe3O4 through an overall reaction [41]: Fe+2 + 2Fe+3 + 8OH-→ Fe3O4.This process is accompanied by release of volatile gases such as HCl, water, and carbon dioxide.FTIR spectra of EC and Fe2O3-EC before and after Zn+2 sorption are shown in Figure 1. EC spectrum showed a prominent peak at 1385 cm−1 which is absent in Fe2O3-EC. A peak at 1028 cm−1 was observed on Fe2O3-EC spectrum which is absent on EC. A strong peak at 561 cm−1 was attributed to Fe-O lattice vibrations of Fe3O4 [36, 39, 42, 43] confirming loading of iron oxide onto the biochar. The OH vibrations were observed at 3300-3500 cm−1 while peaks at around 1590-1614 cm−1 were attributed to aromatic C=C and C=O vibrations. The band at 1385 cm−1 is characteristic symmetric CO stretching vibrations. Upon Zn+2 adsorption the peaks on EC at 3386 cm−1 (νO-H) and 1590 cm−1 (νC=O) shifted to 3423 cm−1 and 1603 cm−1, respectively. There was disappearance of peak at 1385 cm−1 on EC and emergency of a new peak at 1109 cm−1 upon Zn+2 adsorption. The peak at 1028 cm−1 disappeared and a shift in peak from 561 to 595 cm−1 was observed on loaded Fe2O3-EC. Emergence of a new peak at 1110 cm−1 was observed on Zn+2 loaded EC. The FTIR spectra for the Zn+2 or Cu+2 loaded onto the biosorbents were almost similar showing that similar functional groups might have been involved in biosorption.Figure 1
FTIR spectrum for adsorbed and unadsorbed biochars: (a) EC, (b) Fe2O3EC.
(a) (b)The magnetisation characteristics of EC and Fe2O3-EC were done and responses are shown in Figure 2. EC showed no magnetic response while Fe2O3-EC was strongly attracted to the magnet. After submerging Fe2O3-EC in water for 2 weeks, the biosorbent retained excellent magnetic properties. The stability of Fe2O3-EC was tested by immersing samples in highly acidic and basic media at pH 2 and 12, respectively for 2 days. Subjecting the supernatant to the magnet showed no detectable ions confirming that iron was strongly bound in the biosorbent matrix.Figure 2
Magnetic response of Fe2O3-EC.SEM-EDX of Fe2O3-EC is shown in Figure 3. The EDS spectrum confirmed the predominance of iron and oxygen, accounting for more than 70% (atomic ratio) of the total surficial elements. Carbon was also detected on the adsorbent surface by EDS. Fe2O3-EC exhibited an irregular rough surface that can provide binding sites for metal ions.Figure 3
SEM-EDX of Fe2O3-EC.


### 3.2. Batch Adsorption Studies

#### 3.2.1. Effect of pH

Solution pH is one crucial parameter that has a bearing on surface ionization of adsorbent functional groups and metal ion speciation [44, 45]. The effect of pH on Zn+2 and Cu+2 removal is shown in Figure 4. The maximum metal ion adsorption was observed at pH 6. Higher values for adsorption were observed using Fe2O3-EC. Incorporation of iron oxide particles in the biochar introduces more sorption sites via electrostatic interactions [46]. Maximum adsorption efficiencies on Fe2O3-EC were 92% and 79.6% for Cu+2 and Zn+2, respectively. EC attained maximum adsorption efficiencies of 66.1% and 78.8% for Cu+2 and Zn+2, respectively. At low pH, hydroxonium and metal ions compete for adsorption sites [47]. The biosorbent surface becomes more negative with increasing pH making it more readily available for metal ion adsorption. Beyond optimum pH a decrease in metal ion adsorption was observed due to formation of soluble metal ion complexes.Figure 4
Effect of pH on metal ion removal using (a) EC and (b) Fe2O3-EC.
(a) (b)


#### 3.2.2. Effect of Contact Time

The effect of contact time on metal ion adsorption is shown in Figure5. The metal ion removal efficient was high during initial stages. This could be attributed to adsorption of metal ions onto the exterior surface of the biochar and the availability of more vacant surface binding sites [48]. This was followed by gradual metal ion uptake due to ions diffusing into the interior surface before levelling off beyond 65 min as a result of saturation and repulsion between the adsorbed species and the bulk phases [49, 50]. Fe2O3-EC gave a maximum adsorption efficiency of 79.5% and 91.8% for Zn+2 and Cu+2, respectively. EC exhibited longer saturation contact time of more than 100 min giving maximum adsorption efficiency for Zn+2 and Cu+2 of 74.1% and 66.2%, respectively.Figure 5
Effect of contact time on metal ion removal using (a) EC and (b) Fe2O3-EC.
(a) (b)


#### 3.2.3. Effect of Initial Metal Ion Concentration

Figure6 shows the effect of initial metal ion concentration on adsorption. A decrease in adsorption efficiency was observed with increasing initial metal ion concentration. At low metal ion concentration most of the metal ions in solution are adsorbed onto vacant active sites of the biosorbent resulting in significantly high metal ion adsorption efficiency [9]. As the initial metal ion concentration increases active sites become saturated leaving most metal ions in solution. Similar trends were previously reported using other biosorbents [17].Figure 6
Effect of initial metal ion concentration on pollutant removal using (a) EC and (b) Fe2O3-EC.
(a) (b)


#### 3.2.4. Effect of Adsorbent Dosage

According to Sari and Tuzen [51], biosorbent dosage determines the pollutant removal capacity for a given initial concentration. The effect of biosorbent dosage on metal ion removal is shown in Figure 7. The metal ion removal efficient increased with increasing biomass dosage until it levelled off at 1.2 g. Maximum metal ion removal of 79.9% and 86.0% for Zn+2 and Cu+2, respectively, was attained using Fe2O3-EC. EC biochar attained a maximum adsorption efficiency of 76.9% and 72.9% for Zn+2 and Cu+2, respectively. An increase in metal adsorption with increasing biomass is attributed to enhanced biosorption sites. A constant percentage for metal ion removal beyond 1.2 g was due to overcrowding of biosorbent particles as a result of excessive biochar dosage leading to overlapping of adsorption sites.Figure 7
Effect of biosorbent dosage on metal ion removal using (a) EC and (b) Fe2O3-EC.
(a) (b)


#### 3.2.5. Effect of Temperature

The effect of solution temperature on metal ion adsorption on the biosorbents is shown in Figure8. The optimum temperature for metal ion adsorption was in the 25-30°C range. As the temperature was raised beyond 30°C metal ion sorption decreased. This could be attributed to degrading stability of adsorption sites with increasing temperature. High agglomeration rates experienced by EC at elevated temperature drastically reduced its sorption efficiency. The ability to withstand harsh temperature conditions by the iron oxide particles contributed to the resilience demonstrated by Fe2O3-EC.Figure 8
Effect of solution temperature on metal ion removal using (a) EC and (b) Fe2O3-EC.
(a) (b)


### 3.3. Adsorption Isotherms

Adsorption isotherms play a fundamental role in predicting the nature of adsorbate-adsorbent interactions. Such studies help in designing adsorption systems for industrial applications since they provide data for maximum adsorption capacities of adsorbents [52]. Freundlich [53] and Langmuir [54] isotherms were applied in this study to interpret experimental data. The linearized form of the Langmuir isotherm (Figure 9) is represented by the following equation: (3)1qe=1KLqmaxCe+1qmaxwhere qmax is the maximum adsorption capacity (mg g−1), Ce (mg L−1) is the equilibrium concentration, and KL is the Langmuir sorption constant (L mg−1) related to the free energy of sorption. A plot of 1/qe versus 1/Ce produces a straight line with intercept 1/qmax and gradient 1/KLqmax can be used to calculate qmax and b.Figure 9
Langmuir isotherm for Zn+2 and Cu+2 biosorption on Fe2O3-EC.Linearized form for Freundlich isotherm is represented as(4)log⁡qe=ln⁡KF+1nln⁡CeA plot of ln⁡qe versus ln⁡Ce produces a straight line which is used to calculate Freundlich constants KF and n from the intercept and slope, respectively.Calculated values of Langmuir and Freundlich parameters for metal ion adsorption on EC and Fe2O3-EC are shown in Table 2. Based on R2 values presented, metal ion adsorption conformed more to the Langmuir isotherm. Maximum adsorption capacities for Zn+2 were higher than that of Cu+2 perhaps due to lower ionic radius of Zn2+ (0.71 A°) than Cu2+ (0.73 A°). The higher charge density of Zn2+ enhances the attraction for the biosorbent. The constant n measures the deviation of biosorption from linearity (intensity of biosorption). The values of n greater than 1 for Zn2+ and Cu2+ ions indicate the favorable nature of biosorption [55].Table 2
Langmuir and Freundlich constants for the metal ions adsorption on Fe2O3-EC.
Adsorbent Adsorbate Langmuir Freundlich q m a x K L R2 K F n R2 (mg g−1) (L mg−1)  (mg g−1)  Fe2O3-EC Cu2+ 3.53 5.916 0.998 3.31 1.69 0.987 Zn2+ 9.42 0.314 0.995 3.85 1.12 0.907  EC Cu2+ 2.06 6.83 0.980 2.58 1.99 0.978 Zn2+ 5.99 5.71 0.992 5.89 1.53 0.966


### 3.4. Adsorption Thermodynamics

Thermodynamic parameters such as entropy change (ΔS), enthalpy change (ΔH), and the Gibbs free energy change (∆G) are essential in the determination of industrial applications of adsorbents. ∆Go is a critical thermodynamic parameter that provides adsorbent characteristics on the adsorption process [56]. These thermodynamic parameters were evaluated using the following equations: (5)KT=qeCe(6)ln⁡KT=∆SoR+∆HoRT(7)∆Go=∆Ho-T∆Sowhere R is the ideal gas constant, T is temperature in Kelvins, and KT is the thermodynamic equilibrium constant. The intercept and gradient derived from the Vant Hoff’s (4) plot of ln⁡KT as an inverse function of temperature were used to calculate ∆Ho and ∆So shown in Table 3. The calculated Gibbs free energy change values for the adsorption of Cu2+ and Zn2+ were all negative confirming the feasibility and spontaneity nature of the adsorption process. The negative ∆Ho values indicate the exothermic nature of the biosorption processes. The entropy change values for metal ions adsorption were negative indicating that the adsorption process involves an associative mechanism [57]. Metal ion adsorption leads to order through the formation of an activated complex between the adsorbate and adsorbent. The negative values of ∆So reflected that no significant change occurs in the internal structures of the adsorbent during adsorption.Table 3
Thermodynamic parameters derived from Vant Hoff’s plots.
∆ G o (kJ/mol) ∆ S o (kJ/mol/K) ∆ H o (kJ/mol) Temperature (K) 298.15 303.15 308.15 313.15 318.15 323.15 Adsorbent Adsorbate EC Cu2+ -9.571 -8.044 -6.517 -4.991 -3.464 -1.937 -0.305 -100.608 Zn2+ -9.317 -7.604 -5.891 -4.177 -2.464 -0.751 -0.343 -111.474  Fe2O3-EC Cu2+ -6.370 -8.018 -9.667 -11.315 -12.964 -14.612 -0.330 -91.928 Zn2+ -8.896 -8.569 -8.243 -7.916 -7.589 -7.262 -0.065 -28.392


### 3.5. Adsorption Kinetics

Adsorption kinetics provide essential information for determining the adsorption rate controlling mechanisms. Two frequently used integrated linearized kinetic models: Lagergren and Ho-McKay (8) and (9) [17, 58] were used to analyse the experimental data. The parameters qt and qe (mg g−1) are the respective adsorption capacities at time, t, and equilibrium, k1 (min−1) and k2 (g mg−1 min−1) are the Lagergren’s pseudo 1st- and Ho-McKay’s 2nd-order constants, respectively.(8)log⁡qe-qt=log⁡qe-k1t2.303(9)tqt=1k2qe2+tqeEquation (8) hypothesise an adsorption rate that is proportional to the population of vacant sites whilst (9) is based on the hypothesis that the sorption rate is closely related to the square of the population of vacant sites. The biosorption kinetic parameters are shown in Table 4. The coefficients of determination (R2) for the pseudo-first-order on EC were 0.969 and 0.974 for Zn+2 and Cu+2, respectively. On the other hand, the correlation coefficient (R2) values of the pseudo-second-order kinetic model on EC were 0.984 and 0.982 for Zn+2 and Cu+2 ions, respectively. Higher correlation coefficient values and good agreement between calculated (qe) and experimental (qexp) values indicate more applicability of the pseudo-second-order kinetic model for EC systems. The coefficients of determination for the pseudo-first-order on Fe2O3-EC were 0.971 and 0.598 for Zn+2 and Cu+2, respectively. The correlation coefficient values of the pseudo-second-order kinetic model were 0.998 and 0.993 for Zn+2 and Cu+2, respectively. Higher and good agreement between calculated and experimental values and high correlation coefficient values show that the pseudo-second-order kinetic model conforms to the Fe2O3-EC systems. Such kinetic data show that chemisorption was the rate limiting step in the uptake of the metal ions by the adsorbents.Table 4
First- and second-order kinetic parameters for metal ion adsorption on biochars.
Adsorbent Ion q e x p (mg/g) Pseudo-1st order Pseudo-2nd order q e (mg/g) k1 (min−1) R2 q e (mg/g) k2 (g/mg min−1) R2 EC Cu+2 2.46 2.30 0.045 0.974 2.39 0.18 0.982 Zn+2 6.01 6.62 0.049 0.969 5.87 0.05 0.984  Fe2O3-EC Cu+2 3.25 1.41 0.091 0.598 3.24 0.05 0.998 Zn+2 7.91 7.43 0.35 0.971 8.09 0.01 0.993


### 3.6. Desorption and Reusability Studies

The reusability of adsorbents is important to achieve cost effective water treatment. Desorption studies help in the process design of systems by giving information on mechanisms and recovery of adsorbates from industrial wastewater and the adsorbent. Table5 shows the adsorption efficiency of Fe2O3-EC and EC after five desorption cycles. The maximum desorption was attained at 0.25 M NaOH. The adsorption efficiency of Fe2O3-EC remained above 80% in all studies indicative of its high reusability capacity. This data suggests that Fe2O3-EC is cost effective in the treatment of water containing low concentrations (less than 15 mg L−1) of Zn+2 and Cu+2.Table 5
Adsorption efficiency of adsorbents after five desorption cycles.
NaOH concentration (M) Adsorption efficiency (%) EC Fe2O3-EC Cu+2 Zn+2 Cu+2 Zn+2 0.10 73 75 87 88  0.25 75 78 92 93  0.50 69 70 81 82


## 3.1. Characterisation of the Adsorbent

Table1 shows the specific surface areas of Fe2O3-EC and EC at different pyrolysis temperatures. Surface areas of EC were smaller than those of Fe2O3-EC. This could be attributed to Fe2O3-EC containing considerable proportion of iron oxide particles, which have small surface areas and abundant transitional pores [39]. For both adsorbents, an increase in surface area with increasing temperature was observed due to volatilization of transitional fluxes (secondary reactions) presented by the thermal decomposition of the biomass cellulosic and semicellulosic structures [40]. These secondary reactions become prominent with increasing temperature resulting in continued cracking of the biomass structure forming voids on the surface. However, as temperature was further increased beyond 500°C, there was a notable decrease in surface area due to collapse of the overall mesoporous structure of the adsorbents as a result of volatilization of most of the biomass inorganic binding sites [40].Table 1
Specific surface areas of EC and Fe2O3-EC.
Temperature (°C) Adsorbent Surface area (m2 g−1) 350 EC 150.13 Fe2O3-EC 163.60  450 EC 610.88 Fe2O3-EC 1038.48  600 EC 553.59 Fe2O3-EC 983.75The formation of iron oxides could be explained by the fact that during pyrolysis the ferric chloride underwent drastic dehydration process forming Fe3O4 through an overall reaction [41]: Fe+2 + 2Fe+3 + 8OH-→ Fe3O4.This process is accompanied by release of volatile gases such as HCl, water, and carbon dioxide.FTIR spectra of EC and Fe2O3-EC before and after Zn+2 sorption are shown in Figure 1. EC spectrum showed a prominent peak at 1385 cm−1 which is absent in Fe2O3-EC. A peak at 1028 cm−1 was observed on Fe2O3-EC spectrum which is absent on EC. A strong peak at 561 cm−1 was attributed to Fe-O lattice vibrations of Fe3O4 [36, 39, 42, 43] confirming loading of iron oxide onto the biochar. The OH vibrations were observed at 3300-3500 cm−1 while peaks at around 1590-1614 cm−1 were attributed to aromatic C=C and C=O vibrations. The band at 1385 cm−1 is characteristic symmetric CO stretching vibrations. Upon Zn+2 adsorption the peaks on EC at 3386 cm−1 (νO-H) and 1590 cm−1 (νC=O) shifted to 3423 cm−1 and 1603 cm−1, respectively. There was disappearance of peak at 1385 cm−1 on EC and emergency of a new peak at 1109 cm−1 upon Zn+2 adsorption. The peak at 1028 cm−1 disappeared and a shift in peak from 561 to 595 cm−1 was observed on loaded Fe2O3-EC. Emergence of a new peak at 1110 cm−1 was observed on Zn+2 loaded EC. The FTIR spectra for the Zn+2 or Cu+2 loaded onto the biosorbents were almost similar showing that similar functional groups might have been involved in biosorption.Figure 1
FTIR spectrum for adsorbed and unadsorbed biochars: (a) EC, (b) Fe2O3EC.
(a) (b)The magnetisation characteristics of EC and Fe2O3-EC were done and responses are shown in Figure 2. EC showed no magnetic response while Fe2O3-EC was strongly attracted to the magnet. After submerging Fe2O3-EC in water for 2 weeks, the biosorbent retained excellent magnetic properties. The stability of Fe2O3-EC was tested by immersing samples in highly acidic and basic media at pH 2 and 12, respectively for 2 days. Subjecting the supernatant to the magnet showed no detectable ions confirming that iron was strongly bound in the biosorbent matrix.Figure 2
Magnetic response of Fe2O3-EC.SEM-EDX of Fe2O3-EC is shown in Figure 3. The EDS spectrum confirmed the predominance of iron and oxygen, accounting for more than 70% (atomic ratio) of the total surficial elements. Carbon was also detected on the adsorbent surface by EDS. Fe2O3-EC exhibited an irregular rough surface that can provide binding sites for metal ions.Figure 3
SEM-EDX of Fe2O3-EC.


## 3.2. Batch Adsorption Studies

### 3.2.1. Effect of pH

Solution pH is one crucial parameter that has a bearing on surface ionization of adsorbent functional groups and metal ion speciation [44, 45]. The effect of pH on Zn+2 and Cu+2 removal is shown in Figure 4. The maximum metal ion adsorption was observed at pH 6. Higher values for adsorption were observed using Fe2O3-EC. Incorporation of iron oxide particles in the biochar introduces more sorption sites via electrostatic interactions [46]. Maximum adsorption efficiencies on Fe2O3-EC were 92% and 79.6% for Cu+2 and Zn+2, respectively. EC attained maximum adsorption efficiencies of 66.1% and 78.8% for Cu+2 and Zn+2, respectively. At low pH, hydroxonium and metal ions compete for adsorption sites [47]. The biosorbent surface becomes more negative with increasing pH making it more readily available for metal ion adsorption. Beyond optimum pH a decrease in metal ion adsorption was observed due to formation of soluble metal ion complexes.Figure 4
Effect of pH on metal ion removal using (a) EC and (b) Fe2O3-EC.
(a) (b)


### 3.2.2. Effect of Contact Time

The effect of contact time on metal ion adsorption is shown in Figure5. The metal ion removal efficient was high during initial stages. This could be attributed to adsorption of metal ions onto the exterior surface of the biochar and the availability of more vacant surface binding sites [48]. This was followed by gradual metal ion uptake due to ions diffusing into the interior surface before levelling off beyond 65 min as a result of saturation and repulsion between the adsorbed species and the bulk phases [49, 50]. Fe2O3-EC gave a maximum adsorption efficiency of 79.5% and 91.8% for Zn+2 and Cu+2, respectively. EC exhibited longer saturation contact time of more than 100 min giving maximum adsorption efficiency for Zn+2 and Cu+2 of 74.1% and 66.2%, respectively.Figure 5
Effect of contact time on metal ion removal using (a) EC and (b) Fe2O3-EC.
(a) (b)


### 3.2.3. Effect of Initial Metal Ion Concentration

Figure6 shows the effect of initial metal ion concentration on adsorption. A decrease in adsorption efficiency was observed with increasing initial metal ion concentration. At low metal ion concentration most of the metal ions in solution are adsorbed onto vacant active sites of the biosorbent resulting in significantly high metal ion adsorption efficiency [9]. As the initial metal ion concentration increases active sites become saturated leaving most metal ions in solution. Similar trends were previously reported using other biosorbents [17].Figure 6
Effect of initial metal ion concentration on pollutant removal using (a) EC and (b) Fe2O3-EC.
(a) (b)


### 3.2.4. Effect of Adsorbent Dosage

According to Sari and Tuzen [51], biosorbent dosage determines the pollutant removal capacity for a given initial concentration. The effect of biosorbent dosage on metal ion removal is shown in Figure 7. The metal ion removal efficient increased with increasing biomass dosage until it levelled off at 1.2 g. Maximum metal ion removal of 79.9% and 86.0% for Zn+2 and Cu+2, respectively, was attained using Fe2O3-EC. EC biochar attained a maximum adsorption efficiency of 76.9% and 72.9% for Zn+2 and Cu+2, respectively. An increase in metal adsorption with increasing biomass is attributed to enhanced biosorption sites. A constant percentage for metal ion removal beyond 1.2 g was due to overcrowding of biosorbent particles as a result of excessive biochar dosage leading to overlapping of adsorption sites.Figure 7
Effect of biosorbent dosage on metal ion removal using (a) EC and (b) Fe2O3-EC.
(a) (b)


### 3.2.5. Effect of Temperature

The effect of solution temperature on metal ion adsorption on the biosorbents is shown in Figure8. The optimum temperature for metal ion adsorption was in the 25-30°C range. As the temperature was raised beyond 30°C metal ion sorption decreased. This could be attributed to degrading stability of adsorption sites with increasing temperature. High agglomeration rates experienced by EC at elevated temperature drastically reduced its sorption efficiency. The ability to withstand harsh temperature conditions by the iron oxide particles contributed to the resilience demonstrated by Fe2O3-EC.Figure 8
Effect of solution temperature on metal ion removal using (a) EC and (b) Fe2O3-EC.
(a) (b)


## 3.2.1. Effect of pH

Solution pH is one crucial parameter that has a bearing on surface ionization of adsorbent functional groups and metal ion speciation [44, 45]. The effect of pH on Zn+2 and Cu+2 removal is shown in Figure 4. The maximum metal ion adsorption was observed at pH 6. Higher values for adsorption were observed using Fe2O3-EC. Incorporation of iron oxide particles in the biochar introduces more sorption sites via electrostatic interactions [46]. Maximum adsorption efficiencies on Fe2O3-EC were 92% and 79.6% for Cu+2 and Zn+2, respectively. EC attained maximum adsorption efficiencies of 66.1% and 78.8% for Cu+2 and Zn+2, respectively. At low pH, hydroxonium and metal ions compete for adsorption sites [47]. The biosorbent surface becomes more negative with increasing pH making it more readily available for metal ion adsorption. Beyond optimum pH a decrease in metal ion adsorption was observed due to formation of soluble metal ion complexes.Figure 4
Effect of pH on metal ion removal using (a) EC and (b) Fe2O3-EC.
(a) (b)


## 3.2.2. Effect of Contact Time

The effect of contact time on metal ion adsorption is shown in Figure5. The metal ion removal efficient was high during initial stages. This could be attributed to adsorption of metal ions onto the exterior surface of the biochar and the availability of more vacant surface binding sites [48]. This was followed by gradual metal ion uptake due to ions diffusing into the interior surface before levelling off beyond 65 min as a result of saturation and repulsion between the adsorbed species and the bulk phases [49, 50]. Fe2O3-EC gave a maximum adsorption efficiency of 79.5% and 91.8% for Zn+2 and Cu+2, respectively. EC exhibited longer saturation contact time of more than 100 min giving maximum adsorption efficiency for Zn+2 and Cu+2 of 74.1% and 66.2%, respectively.Figure 5
Effect of contact time on metal ion removal using (a) EC and (b) Fe2O3-EC.
(a) (b)


## 3.2.3. Effect of Initial Metal Ion Concentration

Figure6 shows the effect of initial metal ion concentration on adsorption. A decrease in adsorption efficiency was observed with increasing initial metal ion concentration. At low metal ion concentration most of the metal ions in solution are adsorbed onto vacant active sites of the biosorbent resulting in significantly high metal ion adsorption efficiency [9]. As the initial metal ion concentration increases active sites become saturated leaving most metal ions in solution. Similar trends were previously reported using other biosorbents [17].Figure 6
Effect of initial metal ion concentration on pollutant removal using (a) EC and (b) Fe2O3-EC.
(a) (b)


## 3.2.4. Effect of Adsorbent Dosage

According to Sari and Tuzen [51], biosorbent dosage determines the pollutant removal capacity for a given initial concentration. The effect of biosorbent dosage on metal ion removal is shown in Figure 7. The metal ion removal efficient increased with increasing biomass dosage until it levelled off at 1.2 g. Maximum metal ion removal of 79.9% and 86.0% for Zn+2 and Cu+2, respectively, was attained using Fe2O3-EC. EC biochar attained a maximum adsorption efficiency of 76.9% and 72.9% for Zn+2 and Cu+2, respectively. An increase in metal adsorption with increasing biomass is attributed to enhanced biosorption sites. A constant percentage for metal ion removal beyond 1.2 g was due to overcrowding of biosorbent particles as a result of excessive biochar dosage leading to overlapping of adsorption sites.Figure 7
Effect of biosorbent dosage on metal ion removal using (a) EC and (b) Fe2O3-EC.
(a) (b)


## 3.2.5. Effect of Temperature

The effect of solution temperature on metal ion adsorption on the biosorbents is shown in Figure8. The optimum temperature for metal ion adsorption was in the 25-30°C range. As the temperature was raised beyond 30°C metal ion sorption decreased. This could be attributed to degrading stability of adsorption sites with increasing temperature. High agglomeration rates experienced by EC at elevated temperature drastically reduced its sorption efficiency. The ability to withstand harsh temperature conditions by the iron oxide particles contributed to the resilience demonstrated by Fe2O3-EC.Figure 8
Effect of solution temperature on metal ion removal using (a) EC and (b) Fe2O3-EC.
(a) (b)


## 3.3. Adsorption Isotherms

Adsorption isotherms play a fundamental role in predicting the nature of adsorbate-adsorbent interactions. Such studies help in designing adsorption systems for industrial applications since they provide data for maximum adsorption capacities of adsorbents [52]. Freundlich [53] and Langmuir [54] isotherms were applied in this study to interpret experimental data. The linearized form of the Langmuir isotherm (Figure 9) is represented by the following equation: (3)1qe=1KLqmaxCe+1qmaxwhere qmax is the maximum adsorption capacity (mg g−1), Ce (mg L−1) is the equilibrium concentration, and KL is the Langmuir sorption constant (L mg−1) related to the free energy of sorption. A plot of 1/qe versus 1/Ce produces a straight line with intercept 1/qmax and gradient 1/KLqmax can be used to calculate qmax and b.Figure 9
Langmuir isotherm for Zn+2 and Cu+2 biosorption on Fe2O3-EC.Linearized form for Freundlich isotherm is represented as(4)log⁡qe=ln⁡KF+1nln⁡CeA plot of ln⁡qe versus ln⁡Ce produces a straight line which is used to calculate Freundlich constants KF and n from the intercept and slope, respectively.Calculated values of Langmuir and Freundlich parameters for metal ion adsorption on EC and Fe2O3-EC are shown in Table 2. Based on R2 values presented, metal ion adsorption conformed more to the Langmuir isotherm. Maximum adsorption capacities for Zn+2 were higher than that of Cu+2 perhaps due to lower ionic radius of Zn2+ (0.71 A°) than Cu2+ (0.73 A°). The higher charge density of Zn2+ enhances the attraction for the biosorbent. The constant n measures the deviation of biosorption from linearity (intensity of biosorption). The values of n greater than 1 for Zn2+ and Cu2+ ions indicate the favorable nature of biosorption [55].Table 2
Langmuir and Freundlich constants for the metal ions adsorption on Fe2O3-EC.
Adsorbent Adsorbate Langmuir Freundlich q m a x K L R2 K F n R2 (mg g−1) (L mg−1)  (mg g−1)  Fe2O3-EC Cu2+ 3.53 5.916 0.998 3.31 1.69 0.987 Zn2+ 9.42 0.314 0.995 3.85 1.12 0.907  EC Cu2+ 2.06 6.83 0.980 2.58 1.99 0.978 Zn2+ 5.99 5.71 0.992 5.89 1.53 0.966


## 3.4. Adsorption Thermodynamics

Thermodynamic parameters such as entropy change (ΔS), enthalpy change (ΔH), and the Gibbs free energy change (∆G) are essential in the determination of industrial applications of adsorbents. ∆Go is a critical thermodynamic parameter that provides adsorbent characteristics on the adsorption process [56]. These thermodynamic parameters were evaluated using the following equations: (5)KT=qeCe(6)ln⁡KT=∆SoR+∆HoRT(7)∆Go=∆Ho-T∆Sowhere R is the ideal gas constant, T is temperature in Kelvins, and KT is the thermodynamic equilibrium constant. The intercept and gradient derived from the Vant Hoff’s (4) plot of ln⁡KT as an inverse function of temperature were used to calculate ∆Ho and ∆So shown in Table 3. The calculated Gibbs free energy change values for the adsorption of Cu2+ and Zn2+ were all negative confirming the feasibility and spontaneity nature of the adsorption process. The negative ∆Ho values indicate the exothermic nature of the biosorption processes. The entropy change values for metal ions adsorption were negative indicating that the adsorption process involves an associative mechanism [57]. Metal ion adsorption leads to order through the formation of an activated complex between the adsorbate and adsorbent. The negative values of ∆So reflected that no significant change occurs in the internal structures of the adsorbent during adsorption.Table 3
Thermodynamic parameters derived from Vant Hoff’s plots.
∆ G o (kJ/mol) ∆ S o (kJ/mol/K) ∆ H o (kJ/mol) Temperature (K) 298.15 303.15 308.15 313.15 318.15 323.15 Adsorbent Adsorbate EC Cu2+ -9.571 -8.044 -6.517 -4.991 -3.464 -1.937 -0.305 -100.608 Zn2+ -9.317 -7.604 -5.891 -4.177 -2.464 -0.751 -0.343 -111.474  Fe2O3-EC Cu2+ -6.370 -8.018 -9.667 -11.315 -12.964 -14.612 -0.330 -91.928 Zn2+ -8.896 -8.569 -8.243 -7.916 -7.589 -7.262 -0.065 -28.392


## 3.5. Adsorption Kinetics

Adsorption kinetics provide essential information for determining the adsorption rate controlling mechanisms. Two frequently used integrated linearized kinetic models: Lagergren and Ho-McKay (8) and (9) [17, 58] were used to analyse the experimental data. The parameters qt and qe (mg g−1) are the respective adsorption capacities at time, t, and equilibrium, k1 (min−1) and k2 (g mg−1 min−1) are the Lagergren’s pseudo 1st- and Ho-McKay’s 2nd-order constants, respectively.(8)log⁡qe-qt=log⁡qe-k1t2.303(9)tqt=1k2qe2+tqeEquation (8) hypothesise an adsorption rate that is proportional to the population of vacant sites whilst (9) is based on the hypothesis that the sorption rate is closely related to the square of the population of vacant sites. The biosorption kinetic parameters are shown in Table 4. The coefficients of determination (R2) for the pseudo-first-order on EC were 0.969 and 0.974 for Zn+2 and Cu+2, respectively. On the other hand, the correlation coefficient (R2) values of the pseudo-second-order kinetic model on EC were 0.984 and 0.982 for Zn+2 and Cu+2 ions, respectively. Higher correlation coefficient values and good agreement between calculated (qe) and experimental (qexp) values indicate more applicability of the pseudo-second-order kinetic model for EC systems. The coefficients of determination for the pseudo-first-order on Fe2O3-EC were 0.971 and 0.598 for Zn+2 and Cu+2, respectively. The correlation coefficient values of the pseudo-second-order kinetic model were 0.998 and 0.993 for Zn+2 and Cu+2, respectively. Higher and good agreement between calculated and experimental values and high correlation coefficient values show that the pseudo-second-order kinetic model conforms to the Fe2O3-EC systems. Such kinetic data show that chemisorption was the rate limiting step in the uptake of the metal ions by the adsorbents.Table 4
First- and second-order kinetic parameters for metal ion adsorption on biochars.
Adsorbent Ion q e x p (mg/g) Pseudo-1st order Pseudo-2nd order q e (mg/g) k1 (min−1) R2 q e (mg/g) k2 (g/mg min−1) R2 EC Cu+2 2.46 2.30 0.045 0.974 2.39 0.18 0.982 Zn+2 6.01 6.62 0.049 0.969 5.87 0.05 0.984  Fe2O3-EC Cu+2 3.25 1.41 0.091 0.598 3.24 0.05 0.998 Zn+2 7.91 7.43 0.35 0.971 8.09 0.01 0.993


## 3.6. Desorption and Reusability Studies

The reusability of adsorbents is important to achieve cost effective water treatment. Desorption studies help in the process design of systems by giving information on mechanisms and recovery of adsorbates from industrial wastewater and the adsorbent. Table5 shows the adsorption efficiency of Fe2O3-EC and EC after five desorption cycles. The maximum desorption was attained at 0.25 M NaOH. The adsorption efficiency of Fe2O3-EC remained above 80% in all studies indicative of its high reusability capacity. This data suggests that Fe2O3-EC is cost effective in the treatment of water containing low concentrations (less than 15 mg L−1) of Zn+2 and Cu+2.Table 5
Adsorption efficiency of adsorbents after five desorption cycles.
NaOH concentration (M) Adsorption efficiency (%) EC Fe2O3-EC Cu+2 Zn+2 Cu+2 Zn+2 0.10 73 75 87 88  0.25 75 78 92 93  0.50 69 70 81 82


## 4. Conclusion

In this study Fe2O3-EC prepared by coprecipitation of ferric and ferrous salts on water hyacinth powder followed by slow pyrolysis has demonstrated great potential in the removal of aqueous metal ion pollutants. This approach provided a useful way to manage this highly problematic invasive species. Metal ion adsorption onto the biosorbents conformed to the Langmuir isotherm. Chemical modification of hyacinth biochar significantly improved metal ion adsorption capacity. The metal ions adsorption onto the adsorbents best fit pseudo 2nd-order kinetics. Thermodynamic parameters showed that the biosorption process was feasible, spontaneous, and exothermic. Fe2O3-EC exhibited metal ion adsorption efficiency above 80% in all studies indicative of its high reusability capacity.This study focused on sorption of single metal ions from spiked solution. There is need to assess the competitive sorption of metal ions by hyacinth biochar in natural waters. Wastewater is more complicated than simulated water used in this study. To ensure suitability of hyacinth biochar in treating metal ions in wastewater, the use of actual contaminated water for studies should be done. To further support field application, future studies should address factors related to metal ion removal efficiency such as application rate, modification of biochar using other minerals, dosing, regeneration, and disposal of metal ions adsorbed biochars.



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