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Chemical Engineering Communications

ISSN: 0098-6445 (Print) 1563-5201 (Online) Journal homepage: http://www.tandfonline.com/loi/gcec20

STUDY OF HG(II) REMOVAL FROM AQUEOUS SOLUTION USING LIGNOCELLULOSIC COCONUT FIBER BIOSORBENTS: EQUILIBRIUM AND KINETIC EVALUATION K. Johari , N. Saman , S. T. Song , J. Y. Y. Heng & H. Mat To cite this article: K. Johari , N. Saman , S. T. Song , J. Y. Y. Heng & H. Mat (2014) STUDY OF HG(II) REMOVAL FROM AQUEOUS SOLUTION USING LIGNOCELLULOSIC COCONUT FIBER BIOSORBENTS: EQUILIBRIUM AND KINETIC EVALUATION, Chemical Engineering Communications, 201:9, 1198-1220, DOI: 10.1080/00986445.2013.806311 To link to this article: http://dx.doi.org/10.1080/00986445.2013.806311

Accepted author version posted online: 09 Apr 2014. Published online: 09 Apr 2014. Submit your article to this journal

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Date: 05 October 2016, At: 20:18

Chem. Eng. Comm., 201:1198–1220, 2014 Copyright # Taylor & Francis Group, LLC ISSN: 0098-6445 print=1563-5201 online DOI: 10.1080/00986445.2013.806311

Study of Hg(II) Removal From Aqueous Solution Using Lignocellulosic Coconut Fiber Biosorbents: Equilibrium and Kinetic Evaluation K. JOHARI,1 N. SAMAN,1 S. T. SONG,1 J. Y. Y. HENG,2 AND H. MAT1,3 1

Advanced Materials and Process Engineering Laboratory, Faculty of Chemical Engineering, Universiti Teknologi Malaysia, Skudai, Johor, Malaysia 2 Department of Chemical Engineering, Imperial College London, London, UK 3 Novel Materials Research Group, Nanotechnology Research Alliance, Universiti Teknologi Malaysia, Skudai, Johor, Malaysia Lignocellulosic coconut wastes such as pith and fiber, which are abundantly available and cheap, have the potential of being used as low-cost biosorbents for heavy metal ion removal. In this study, pristine (CF-Pristine) and NaOH-treated (CF-NaOH) coconut fibers were used as a biosorbent for Hg(II) removal from an aqueous solution. The coconut fiber biosorbent (CFB) was characterized by scanning electron microscopy (SEM) and Fourier transform-infrared (FTIR) spectroscopy. The Hg(II) sorption capacities obtained for CF-Pristine and CFNaOH were 144.4 and 135.0 mg=g, respectively. Both the equilibrium and kinetic data of Hg(II) sorption onto CFB followed the Langmuir isotherm model and a pseudo-second-order kinetic model, respectively. A further analysis of the kinetic data suggested that the Hg(II) sorption process was governed by both intraparticle and external mass transfer processes, in which film diffusion was the rate-limiting step. These results demonstrated that both pristine- and alkali-treated coconut wastes could be potential low-cost biosorbent alternatives for the removal of Hg(II) from aqueous solutions, such as water containing Hg(II) produced in the oil and gas industry. Keywords Biosorbents; Coconut fiber; Mercury ion; Modification; Sorption

Introduction Mercury is one of the most toxic heavy metals with the potential to contaminate the environment and the ability to accumulate in animals and plants (Wang et al., 2009) and enter the human body through the food chain, causing damage to the human central nervous system (Rao et al., 2009). Mercury may exist not only as a metal but also as inorganically or organically bound mercury (Kraepiel et al., 2003; Matlock et al., 2001; Risher, 2003). The major sources of mercury pollution to the Address correspondence to H. Mat, Advanced Materials and Process Engineering Laboratory, Faculty of Chemical Engineering, Universiti Teknologi Malaysia, 81310 UTM Skudai, Johor, Malaysia. E-mail: [email protected]

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environment are waste effluents from the metal plating industry, mining operations, fertilizer industry, tanneries, and textile industry (Wu et al., 2007). Several methods have been developed for mercury ion removal from wastewater streams, including processes such as ion exchange, chemical precipitation, membrane filtration, adsorption, and photo reduction (Anirudhan et al., 2008; Chowdhury et al., 2011; Ramachandra et al., 2008). As a result, the search for low-cost adsorbents, such as ones made from biological waste materials, has gained considerable academic attention in the past two decades (Anirudhan et al., 2008; Bhatnagar et al., 2010; Demirbas et al., 2002; Kavitha and Namasivayam, 2007; Mahvi, 2008; Phan et al., 2006; Pollard et al., 1992; Shareef, 2009). Common agricultural wastes (AW) such as oil palm (Elaeis guineensis), rice (Oryza sativa L.), and coconut (Cocos nucifera L.) residues are reported to have a high adsorption capacity for heavy metals from wastewater (Anirudhan et al., 2008; Bhatnagar et al., 2010; Chowdhury et al., 2011; Kavitha and Namasivayam, 2007; Moreno-Piraja´n et al., 2011; Phan et al., 2006; Pollard et al., 1992; Wong et al., 2003). The advantages of using AWs as biosorbents include low cost, high sorption efficiency, the minimization of chemical or biological sludge production, regeneration potential of biosorbents, and possibility of metal recovery (Suryavanshi and Shukla, 2009). Among the numerous AWs, coconut wastes, such as coconut husks (fiber and pith) and coconut shells, have been extensively investigated as biosorbents for the removal of diverse heavy metal ions (Anirudhan et al., 2008; Geethammaa et al., 1998; Hasany and Ahmad, 2006; Sreedhar and Anirudhan, 2000), dyes (Chowdhury et al., 2011; Jain and Shrivastava, 2008; Tan et al., 2008), inorganic anions (Namasivayam and Sangeetha, 2004), radionuclides (Parab and Sudersanan, 2010), and miscellaneous pollutants (Igwe et al., 2008) from water. Only a few studies have investigated the use of coconut husk as biosorbent for the removal of mercury ions from aqueous solution (Hasany and Ahmad, 2006; Sreedhar and Anirudhan, 2000). The sorption of Hg(II) on poly(hydroxyethylmethacrylate) grafted onto coir pith has also been reported (Anirudhan et al., 2007, 2008). The feasibility of employing coconut husk (pith and fiber) as a biosorbent is due to its low price and availability in tropical countries has been reported before (Bismark et al., 2001). It is constituted of (dry weight percentage) holocellulose (70.4%), a-cellulose (45.0%), hemicellulose (30.0%), lignin (25.5%), ash (2.2%), and extractives (2.2%) (Orlando et al., 2003). The coconut fiber (or coir fiber; CF) is extracted from the fibrous husk (mesocarp), which constitutes about 25 wt.% of the nuts (Tomczak et al., 2007). CF comprises (weight percentage) cellulose (36–43%), hemicellulose (41–45%), lignin (0.15–0.25%), and pectin (3–4%) (Gu, 2009; Rahman and Khan, 2007). It is not brittle like glass fibers, amenable to chemical modification, possesses no waste disposal problems since it is nontoxic, possesses high weather resistance due to its high lignin content, absorbs less water than other fibers because of its lower cellulose content, and is basically cheaper than sisal and jute fibers (Geethammaa et al., 1998; Rahman and Khan, 2007). The use of CF as a biosorbent may require surface modification to enhance its metal ion sorption capacity and selectivity. Several treatment methods are known for natural fibers, including alkali treatment (David et al., 1988; Li et al., 2007; Sreenivasan et al., 1996), esterification (Li et al., 2007; Wong et al., 2003), silane treatment (Li et al., 2007), and acetylation (Li et al., 2007; Sreekala et al., 1997). David et al. (1988) reported that the chemical treatment of cellulosic materials

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could lead to changes in the physical and chemical structures of the fiber surfaces. Igwe et al. (2008) modified CF through a thiolation process and successfully used the product as a sorbent for the removal of Hg, As, and Pb ions from an industrial effluent. They found that the modified CF had a lower metal sorption capacity than unmodified CF. Other studies included the use of chemically modified and functionalized CF as biosorbents for the removal of heavy metal ions such as Ni(II), Zn(II), Fe(II), As(III), and Cd(II) (de Sousa et al., 2010; Shukla et al., 2006). In contrast to findings by Igwe et al. (2008), their results indicated that modified CF possessed a higher sorption capacity than unmodified fiber (pristine) in adsorbing metal ions (Shukla et al., 2006). Even though there is a vast potential for the use of CF as a Hg(II) biosorbent, the investigation of Hg(II) sorption onto either pristine or modified CF is still very much limited. Thus, in this study, CF was modified by alkali (NaOH) treatment (CF-NaOH) and its Hg(II) sorption capacity in an aqueous mercury solution was compared to that of pristine CF (CF-Pristine). The alkaline (NaOH) treatment was chosen since it is a simple and relatively cheap method for modifying surface lignocellulosic materials. The sorption of Hg(II) onto CF biosorbents (CFB) was studied in batch sorption experiments with various initial pHs, contact times, and Hg(II) concentrations, while all other parameters were kept constant. The experimental results were then analyzed in terms of equilibrium and kinetics of Hg(II) sorption behavior. The results of the work are believed to provide further insight into the conceptual potential and feasibility of the development of CF-based biosorbents for industrial applications, such as required in production or treatment of wastewater containing Hg(II) from oil and gas production activities.

Materials and Methods Materials CF was provided by a local company (T&H Coconut Fiber Sdn. Bhd.). Analyticalgrade chemicals such as sodium hydroxide (99%), nitric acid (36.5–38.0%), and mercury nitrate monohydrate (Hg(NO3)2) were obtained from Merck (Germany) and J.T. Baker (Europe) respectively. All chemicals were used as received. The deionized water was prepared by using a Purite Water System (UK) model Analyst HP 40. CFB Preparation The CF was ground and sieved to obtain a particle size of 0.3–0.5 mm. As a washing step, the particles were then dispersed into deionized water and allowed to sediment. The supernatant containing foreign surface-adhered particles and water-soluble materials was removed. This step was repeated several times, followed by a drying step in a convection oven at 50.0  0.5 C. This CF biosorbent was denoted CFPristine. The alkali treatment, known as mercerization, was carried out by leaving CF-Pristine in a 5% NaOH solution for 24 h, followed by further washing using deionized water and acetic acid (Sreekala et al., 1997). The sample was then dried in oven at 50.0  0.5 C. The NaOH-treated CF-Pristine was denoted CF-NaOH. CFB Characterizations The surface morphology of the CFB (CF-Pristine and CF-NaOH) was analyzed by scanning electron microscopy (SEM) (JEOL model JSM-6390LV). A Fourier

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transform-infrared (FT-IR) spectrophotometer (PerkinElmer Model 2000) was used to determine the existence of functional groups on the CFB surfaces. The measurements were carried out over 4000–400 cm1 using the KBr disk method. Batch Sorption Experiments The sorption experiments were performed according to the batch sorption method. Hg(II) stock solutions were prepared for experiments using Hg(NO3)2 salt. In a typical equilibrium sorption experiment, 25 mg CFB was suspended in 25 mL of aqueous Hg(II) solution at a concentration of 50 mg=dm3. The pH of the suspension was adjusted to the desired values (pHi 2–10) using HNO3 and NaOH solutions. The mixture solution was shaken at 200 rpm using a temperature-controlled shaker for two days at 30.0  0.5 C. Two days of shaking time were sufficiently long for the suspension to achieve equilibrium conditions. In a subsequent step, the suspension was filtered using a nylon membrane filter (0.80 mm) in order to remove the CFB particles from the solution that could interfere in the mercury concentration determination, using a flame atomic absorption spectrophotometer (AAS) (PerkinElmer model Precisely HGA 900). The effect of Hg(II) concentration (25 to 200 mg=dm3) was determined according to the same procedure as above. The time dependence of the sorption phenomenon was also analyzed according to the protocol of the equilibrium sorption experiment, but using 200 mL of 50 mg=dm3 Hg(II) solution having the same CFB-to-liquid ratio as that of the equilibrium sorption experiment. The suspension was stirred at 200 rpm for up to three days, while 1 mL aliquot was taken at selected times. The aliquot was withdrawn using a glass syringe fitted with a nylon membrane filter (0.80 mm). The Hg(II) concentration was determined using AAS. The amount of Hg(II) sorbed at equilibrium was calculated by qe ¼ ½ðCo  Ce Þ  V =W where qe is the amount of Hg(II) sorbed (mg=g), Co and Ce are the Hg(II) concentrations of the solutions (mg=dm3) before and after equilibrium, respectively, W is the weight of CFB (g), and V is the volume of the Hg(II) solution (mL). In the timedependence sorption study, the amount of Hg(II) sorbed at time t, qt (mg=g), was calculated by qt ¼ ½ðCo  Ct Þ  V =W

Results and Discussions CFB Characterization Figure 1 shows the SEM images of CFB, indicating that the surface morphology of the CF-NaOH was rougher than that of CF-Pristine. In addition, the pores became clear and a large number of regularly placed holes were also observed on the CFNaOH surface that might be due to the leaching out of the waxy cuticle layer (Rout et al., 2001). Rahman and Khan (2007) reported that alkali treatment would led to an increase in surface roughness, the amount of cellulose exposed on the fiber surface, the amount of amorphous cellulose at the expense of crystalline cellulose, and the removal of hydrogen bonding in the network structures. Similar results were

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Figure 1. SEM images of CFB: (a) CF-Pristine, (b) CF-NaOH, (c) CF-Pristine after adsorption, and (d) CF-NaOH after adsorption.

also reported for alkali-treated oil palm fibers (Sreekala et al., 1997). The SEM images of CF-Pristine and CF-NaOH after Hg(II) sorption are shown in Figures 1(c) and 1(d), respectively. It was found that the morphology of samples (c) and (d) changed to rough and broken surfaces after Hg(II) sorption. The FT-IR spectra of CFB in the range of 4000–400 cm1 are shown in Figure 2 and their likely origins are given in Table I. The existence of functional groups such

Figure 2. FT-IR spectra of CFB over 400 to 4000 cm1.

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Table I. FT-IR results of CFB: peak wave numbers, probable peak origins, and peak intensities Wave numbers (cm1) 3438 2919–2892 1734 1638–1624 1505–1470 1261 1038–1036 a

Probable peak origins OH stretching CH, CH2 stretching C=O stretching OH bending C=C stretching vibration C-O-C stretching C-OH stretching vibration

Peak intensitiesa S M W–M M W–M M S

S ¼ strong, M ¼ medium, and W ¼ weak.

as C-O, C=O, C-H, and O-H were the main characteristics attributed to the presence of cellulose, hemicelluloses, and lignin, and thus are the main characteristics of the natural fibers (Rout et al., 2001; Sreekala et al., 1997; Sreenivasan et al., 1996). A broad peak for both CFBs at 3400 cm1 indicated the hydrogen bonded (OH) stretching vibration from the cellulose of CF (Anirudhan et al., 2008). As stated by Sreekala et al. (1997), the mercerization treatment led to an increase of the amount of amorphous cellulose at the expense of crystalline celullose, thereby increasing the FT-IR intensity of the OH peak of the CF-NaOH spectrum. In addition, the typical absorption band at 2900 cm1 indicated the CH stretching vibration of the CH2 from the cellulose and hemicellulose (Rout et al., 2001; Sreenivasan et al., 1996). A peak in the region around 1734 cm1 could be seen for untreated fiber (CF-Pristine) that resulted from the C=O groups, but for the modified fiber (CF-NaOH), it was diminished, which could be the result of NaOH treatment. This might be due to the removal of the reducible hemicellulose during the alkali treatment process (Rout et al., 2001; Sreenivasan et al., 1996). The FT-IR spectra of CF-Pristine and CF-NaOH showed peaks at about 1626 cm1 indicating the peak of OH bending. However, the peak intensity for CF-NaOH increased, indicating an increase of adsorbed water molecules (Sreenivasan et al., 1996). The peak at 1730 cm1 was indicative of C=O stretching vibration, while the band at 1505 cm1 was probably due to C=C stretching vibration. The bands at 1470 cm1 arose from symmetric deformation of the CH2 group of cellulose, while the band at 1272 cm1 was associated with the presence of C-O-C asymmetric stretching. A strong absorption band at 1044 cm1 was related to the C-OH stretching vibration (Mothe and de Miranda, 2009; Rout et al., 2001; Sreenivasan et al., 1996).

Sorption Parameters Effect of pH The pH of the sorbate solution is an important parameter that controls the sorption process (Anirudhan et al., 2008). In this study, the effect of pH on Hg(II) sorption onto CFB was studied at different initial pHs ranging from 2.0 to 10.0, and the results are presented in Figure 3. The changes in pH of the Hg(II) solution showed that the sorption of Hg(II) increased with an increase of pH up to between pH 4.0

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Figure 3. Effect of pH on Hg(II) sorption onto CFB. Experimental conditions: Hg(II) ¼ 50 mg=dm3, contact time ¼ 2 days, temperature ¼ 30  0.5 C, and agitation speed ¼ 200 rpm.

and 6.0, and thereafter it decreased. This could be due to the fact that at low pH values, available H3Oþ ions might compete with Hg(II) ions for active sites on the CFB surfaces, which reduced the total amount of Hg(II) adsorbed (Basha et al., 2009). With the increase of pH, increased rates of Hg(II) sorption were observed due to the decrease of H3Oþ concentration and the formation of positively charged Hg(OH)þ. However, at higher pH values, it was observed that the sorption of mercury decreased as Hg(OH)2 became the dominant species in the solution (Das et al., 2007). Sreedhar and Anirudhan (2000) reported similar results for the sorption of mercury (II) from an aqueous solution onto grafted coconut husk. Effect of Agitation Time Figure 4 shows the effect of agitation time on the sorption of Hg(II) onto CFB. The amount of Hg(II) sorbed increased with an increase of agitation time. This result revealed that the Hg(II) sorption was fast at the initial stages of the contact period, and thereafter it became slow when approaching equilibrium. It was observed that CFB took only about 60 min to achieve 47% Hg(II) sorption, which might be due to the fact that a large number of vacant surface sites were available during the initial stage of the sorption process. After about 120 min, the Hg(II) sorption became slower, which might be due to the reduced number of vacant surface sites and the difficulty of Hg(II) sorbing as a result of the repulsive forces between the Hg(II) in the solid and bulk phases (Kavitha and Namasivayam, 2007). CFB finally reached equilibrium after approximately 6 h. Effect of Initial Hg(II) Concentration The effect of initial mercury concentration on Hg(II) sorption onto CFB is shown in Figure 5. At constant initial pH (pHi ¼ 7.5), increasing the initial concentration of Hg(II) solution from 12 to 350 mg=dm3 resulted in an increase of Hg(II) sorption onto CFB. The results showed that higher initial Hg(II) concentrations could result

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Figure 4. Effect of agitation time on Hg(II) sorption onto CFB. Experimental conditions: Hg(II) ¼ 50 mg=dm3, initial pH ¼ 7.5, temperature ¼ 30  0.5 C, and agitation speed ¼ 200 rpm.

in larger amounts of Hg(II) sorption onto CFB until it reached saturation conditions. Overall, the results showed that CF-Pristine had significantly higher sorption capacity than CF-NaOH, which allowed the conclusion that NaOH treatment of CF did not lead to higher Hg(II) sorption. A similar result reported by Igwe et al. (2008) showed that unmodified coconut fiber was found to have higher Hg(II) sorption capacity than thiolated CF sorbent. Shukla et al. (2006), in contrast, reported that the sorption of Ni(II), Zn(II), and Fe(II) on hydrogen peroxide–modified coconut fibers resulted in higher metal ion uptake than on unmodified coconut fibers. The Hg(II) sorption capacities of CF-Pristine and CF-NaOH were 144.4 and 135.0 mg=g respectively, which are considered to be higher than that of some low-cost adsorbents reported in the literature (Inbaraj et al., 2009; Kadirvelu et al.,

Figure 5. Effect of Hg(II) initial concentration on Hg(II) sorption onto CFB. Experimental conditions: initial pH ¼ 7.5, contact time ¼ 2 days, temperature ¼ 30  0.5 C, and agitation speed ¼ 200 rpm.

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2004; Zabihi et al., 2010). For instance, Sreedhar and Anirudhan (2000) reported that the adsorption of Hg(II) ions onto acrylamide-grafted coconut husk exhibited Hg(II) sorption capacity of 124.7 mg=g at pH 6.0. However, Hg(II) removal from aqueous phase using poly(hydroxyethylmethacrylate) grafted onto coir pith was found to have a sorption capacity of 254.5 mg=g achieved at pH 6.0 (Anirudhan et al., 2007). CFB Sorption Isotherms Several sorption isotherms have been proposed, ranging from single to multiparameter sorption isotherm models (Basha et al., 2008; Foo and Hammed, 2010). In metal ion sorption studies, two-parameter (e.g., Langmuir, Freundlich, Temkin, Halsey, and Dubinin-Radushkevich), three-parameter (e.g., Redlich-Peterson, Sips, and Toth), four-parameter (e.g., Fritz-Schluender), and five-parameter (e.g., FritzSchluender) sorption isotherm models are commonly used to analyze experimentally obtained isotherm data (Basha et al., 2008; Foo and Hammed, 2010). The isotherm data for Hg(II) sorption onto CFB are presented in Figure 6. The Hg(II) sorption equilibrium experiments were carried out at different initial Hg(II) concentrations, ranging from 12 to 350 mg=dm3, the initial pH of 7.5, temperatures of 30.0  0.5 C, and within sorption time of two days. The results showed that Hg(II) sorption increased with an increase of initial mercury concentration (Ci) and thus the equilibrium Hg(II) concentration (Ce). These sorption isotherms were further analyzed using the Langmuir, Freundlich, Dubinin-Radushkevich (DR), and Redlich-Peterson (RP) isotherm models. The Langmuir isotherm model assumes monolayer sorption onto the sorbent surfaces with a finite number of identical sites (i.e., homogeneous surfaces), given by Equation (1): qe ¼ ðqmax KL Ce Þ=ð1 þ KL Ce Þ

ð1Þ

where Ce is the equilibrium of sorbate concentration in the liquid phase (mg=dm3), qe is the amount adsorbed at equilibrium (mg=dm3), and qmax and KL are Langmuir

Figure 6. Isotherms of Hg(II) sorption onto CFB. Experimental conditions: initial pH ¼ 7.5, contact time ¼ 2 days, temperature ¼ 30  0.5 C, and agitation speed ¼ 200 rpm.

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constants related to the monolayer sorption capacity of the sorbent (mg=g) and the free energy of the sorption process (KL ¼ e–DG=RT), respectively. The constant KL denotes the reciprocal of the concentration at which the sorbent is half saturated. The characteristics of the Langmuir isotherm model can be expressed using a dimensionless equilibrium parameter, RL ¼ 1=(1 þ KLCo), where KL is the Langmuir constant and Co is the initial concentration (mg=g). The value of RL indicates the type of isotherm and can be either unfavorable (RL > 1), linear (RL ¼ 1), favorable (0 < RL), or irreversible (RL ¼ 0). The Freundlich isotherm model assumes that sorption occurs on a heterogeneous surface with a nonuniform distribution of heat of sorption over the surface, which can be expressed by: qe ¼ KF  Ce1=n

ð2Þ

where KF and 1=n are the Freundlich constants related to the sorption capacity of the sorbent (dm3=g) and sorption intensity, respectively. The magnitude of 1=n gives a measure of the favorability of sorption, where a value of 1=n less than 1 represents favorable sorption. The Dubinin-Radushkevich (D-R) sorption isotherm model, which is suitable for porous sorbents, is given by Equation (3) and was also used to describe the sorption of metal ions onto biosorbents. qe ¼ qmax expðbe2 Þ

ð3Þ

Here, b is related to the sorption energy (mol2=kJ2) and e is the D-R constant. The constant e ¼ RT ln (1 þ 1=Ce), where R is the gas constant (8.3145 J=mol  K). The constant bpgives the mean free energy of sorption E (kJ=mol) and can be calculated by E ¼ 1= (2  b). Another commonly used isotherm model to analyze metal ion sorption onto biosorbents is the Temkin isotherm model (Basha et al., 2008; Chowdhury et al., 2011; Isik, 2008; Kavitha and Namasivayam, 2007; Lataye et al., 2006). This model assumes a linear decrease of the heat of sorption as opposed to a logarithmic correlation as implied in the Freundlich isotherm model. It can be represented by: qe ¼ RT=bT lnðKT Ce Þ

ð4Þ

Equation (4) can be expressed in its linear form as: qe ¼ B ln KT þ B ln Ce

ð5Þ

where B ¼ RT=bT, R is the gas constant (8.314 J=mol  K), T is the absolute temperature (K), KT is the equilibrium binding constant (dm3=g), and bT is the model constant; B is thus related to the heat of sorption (J=mol). The Redlich-Peterson (R-P) isotherm model, which can be applied to both homogeneous and heterogeneous systems, is also commonly used in analyzing the sorption of heavy metal equilibrium data (Lataye et al., 2006; Pe´rez Marin et al., 2009). It can be described by (Wu et al., 2010): qe ¼ ðKR Ce Þ=ð1 þ aR Ceo Þ

ð6Þ

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where KR (dm3 g1) and aR (dm3 mg1) are the R-P constants, and b is an fitting parameter having values between 1 and 0. For high concentrations, Equation (6) reduces to Equation (2) (Freundlich equation) with KF ¼ KR=aR, and the heterogeneity factor given by 1=n becomes 1-b. For b ¼ 0 and b ¼ 1, Equation (6) reduces to Equation (1) (Langmuir equation) with KL ¼ aRand Equation (7) (Henry’s equation), respectively (Lataye et al., 2006). qe ¼ qmax  KL  Ce

ð7Þ

The constant parameters were calculated through regression using a linear form of the isotherm equations. The parameters and correlation coefficients, R2, are summarized in Table II. Figure 7 shows a comparison of the CFB experimental and the predicted isotherm data obtained from the five isotherm models. The best fitting isotherms for CFB were found to be the Langmuir, Temkin, and Redlich-Peterson (R-P) isotherm models, while the Freundlich and Dubinin-Radushkevich models did not fit well into the experimental isotherm data. The experimental isotherm data were observed fitting well the ‘‘L’’ type curve representing the Langmuir isotherm, where the curve reached a strict asymphotic plateau. This was further supported by comparing the values of R2; the CFB experimental isotherm data had a higher Table II. Langmuir, Freundlich, Dubinin-Radushkevich (D-R), and Redlich-Peterson (R-P) isotherm model parameters and correlation coefficients, R2, for Hg(II) sorption onto CFB at 30  0.5 C CFB Isotherm models Langmuir qmax (mg=g) KL (dm3=mg) R2 Freundlich KF (dm3=g) 1=n R2 Dubinin-Radushkevich (D-R) b  109 (mol2=kJ2) qmax (mg=g) R2 Temkin KT (dm3=g) bT (kJ=mol) R2 Redlich-Peterson (R-P) KR (dm3=g) aR B R2

CF-Pristine

CF-NaOH

166.670 0.033 0.994

142.860 0.042 0.989

9.198 0.540 0.970

11.034 0.471 0.977

4.000 461.156 0.977

5.000 630.650 0.980

125.774 17.373 0.974

168.174 20.992 0.971

5.708 0.116 0.777 0.998

15.070 0.632 0.674 0.973

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Figure 7. Isotherm models analysis of Hg(II) sorption onto CFB. The experimental data are the same as in Figure 5.

R2 for the Langmuir isotherm, which indicated the monolayer sorptions of Hg(II) onto the surface of CFB, with the maximum sorption capacities (qmax) for CFPristine and CF-NaOH of 166.67 and 142.86 mg=g, respectively. The applicability of the Langmuir isotherm model was also reported for the adsorption of Ni(II), Zn(II), and Fe(II) onto coconut fibers (Shukla et al., 2006) and Cd2þ onto coconut fibers functionalized with thiophosphoryl (P=S) group (de Sousa et al., 2010). Hg(II) adsorption onto grafted poly(hydroxyethylmethacrylate) coconut coir pith fitted well into the Langmuir isotherm model (Anirudhan et al., 2007) and, in contrast to their earlier report, into the Freundlich isotherm model (Anirudhan et al., 2008). The analysis of the mean free energy, E (kJ=mol), of the D-R isotherm model can be used to determine whether it is chemical or physical sorption. It was reported that if adsorption energy < 8 kJ=mol, physical sorption would take place, with 8 to 16 kJ=mol corresponding to the chemical ion exchange that dominated the sorption process, and if adsorption energy was 20–40 kJ=mol, the process would be chemical sorption (Itodo and Itodo, 2010; Kara and Demirbel, 2012; Munir et al., 2010; Zolgharnein and Shahmoradi, 2010). In this study, the mean free energy of sorption,

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E, was found to be 11.180 kJ=mol for CF-Pristine and 10.000 kJ=mol for CF-NaOH, which suggests that Hg(II) sorption onto CFB was through the chemical ion-exchange process. In the Temkin isotherm model, the constants related to heat of sorption (bT) were found to be 17.373 and 20.992 kJ=mol for both CFBs, which fell in the range of chemisorption (20–40 kJ=mol). CFB Sorption Kinetics The kinetics of the Hg(II) sorption process onto CFB were analyzed in terms of linear regression of pseudo-first-order kinetics (Lagergren model), pseudo-secondorder kinetics (Ho and McKay model), and the Elovich model. The pseudo-firstorder kinetic model, given by Equation (9), has been widely used to predict metal ion sorption kinetics (Qiu et al., 2009): logðqe  qt Þ ¼ log qe  k1 t

ð9Þ

where qe and qt (mg=g) are the sorbed amount at equilibrium and at time t (min), respectively, and k1 is the Lagergren rate constant of the pseudo-first-order kinetic model (min1). The sorption kinetic data could also be described by using pseudo-second-order kinetic model, given by (Ho and McKay, 1998, 1999; Qiu et al., 2009) t=qt ¼ 1=k2 q2e þ t=qe

ð10Þ

where qe and qt (mg=g) are the sorbed amount at equilibrium and at time t (min), respectively, (mg=g) and k2 is the rate constant of the pseudo-second-order kinetic model (g=mg min). The initial sorption rate, h (mg=g min), at t ¼ 0 was defined as h ¼ k2 qe2 . Equation (11) shows the Elovich model, which describes the sorption process onto heterogeneous surfaces (Pe´rez Marin et al., 2009): qt ¼ ð1=bÞ ln ab þ ð1=bÞ ln t

ð11Þ

where qt is the sorption capacity at time t (mg=g), a is the initial sorption rate (mg=g  min), and b is the Elovich constant (g=mg). Figure 8 shows nonlinear regression fittings of Hg(II) sorption onto CFB using the parameters obtained from the method of linear regression analysis (Table III). The experimental qe values for CF-Pristine and CF-NaOH were 38.50 and 36.00 mg=g, respectively. It was observed that the pseudo-first-order kinetic model resulted in a significant deviation from the experimental kinetic data, and the calculated qe value was lower than the experimental qe value even though the R2 value (>0.963) was found comparable in both the pseudo-second-order kinetic and Elovich models. It was reported that a high R2 from linearization fitting does not necessarily indicate the best fitting model to describe nonlinear experimental data (Badertscher and Pretsch, 2006; El-Khaiary and Malash, 2011). The comparison between the pseudo-second-order kinetic and Elovich models using R2 showed that the pseudo-second-order kinetic model should be the best model to describe the entire Hg(II) sorption process (R2 > 0.983). In addition, the calculated qe obtained from the pseudo-second-order kinetic model equation was in close agreement with the

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Figure 8. Kinetic models analysis of Hg(II) sorption onto CFB. The experimental data are the same as in Figure 4.

experimental qe value. Thus, this indicated that the kinetics of the Hg(II) sorption process could be better described by using the pseudo-second-order kinetic model over the whole range of the sorption process (David et al., 1988; Hameed et al., 2008). Igwe et al. (2008) reported similar results for Hg(II) ion sorption onto both unmodified and thiolated coconut fibers. For Hg(II) sorption onto grafted poly (hydroxyethylmethacrylate), the coconut coir pith kinetic data fitted well to first-order reversible kinetics (Anirudhan et al., 2008), in contrast to the authors’ earlier report that the kinetic data followed the pseudo-second-order kinetic model (Anirudhan et al., 2007). CFB Sorption Mechanism The time dependence of Hg(II) sorption onto CFB was found to be rapid at the initial stage and became slow with an increase of contact time. It was found that the pseudo-second-order kinetic model gave the best fitting to the kinetic data. However, this reaction-based kinetic model does not physically represent the actual sorption process. In a sorption process, it is generally assumed and accepted that the sorption of metal ions onto biosorbents is governed by the migration of metal ions in

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K. Johari et al. Table III. Pseudo-first-order, pseudo-second-order model, and Elovich constants and correlation coefficients, R2, for Hg(II) sorption onto CFB at 30  0.5 C CFB Kinetic models Pseudo-first-order qe (mg=g) k1  103 (min1) R2 Pseudo-second-order qe (mg=g) k1  103 (g=mg  min) h (mg=g  min) R2 Elovich a (mg=g  min) b (g=mg) R2

CF-Pristine

CF-NaOH

26.900 21.000 0.978

24.030 19.000 0.963

40.000 1.500 2.558 0.983

38.460 1.800 2.660 0.988

22.040 0.211 0.916

18.580 0.212 0.945

the bulk phase, diffusion in the boundary film, diffusion in biosorbent pores, and metal ion-biosorbent surface interactions (Garcia-Reyes and Rangel-Mendez, 2010; Kumar and Gaur, 2011; Ofomaja, 2010; Phan et al., 2006). These physical steps can be described by analyzing the process using diffusion-based kinetic models such as the intraparticle diffusion model, the external mass transfer model, and the ¨ nal et al., 2007). Boyd plot (O For Hg(II) sorption onto CFB, the mechanism may be assumed to involve: (1) migration of Hg(II) from the bulk of the solution to the CFB surfaces, (2) diffusion of Hg(II) through the boundary layer to the CFB surfaces (film diffusion), (3) sorption of Hg(II) onto the external and internal CFB surfaces due to the interaction of surface functional groups (e.g., OH) and Hg(II) (where the surfaces are expected to be negatively charged, which could facilitate the sorption of positively charged Hg(II)), and (4) diffusion of Hg(II) into the interior CFB pores. The bulk migration of Hg(II) in the solution phase is commonly considered to be very rapid due to efficient mixing, and, thus, the mass transfer process is predominantly controlled by convection rather than a diffusion process (Kumar and Gaur, 2011). The surface interactions leading to Hg(II) sorption, which include mechanisms such as physicalchemical sorption, ion exchange, precipitation, and complexation, are also considered to be fast processes as reported by many researchers working on heavy metal sorption onto biosorbents (Garcia-Reyes and Rangel-Mendez, 2010; Igwe et al., 2008). This fast reaction may be due to the existence of functional groups (e.g., aliphatic hydroxyl, polyphenolic, ether, or carboxylic) on the surface of the biosorbents. For CFB, the mercerization process was assumed to form more reactive functional groups such as hydroxyl groups on the CFB surfaces. Chowdhury et al. (2011) reported that chemical treatment using sodium hydroxide may remove lignin and other impurities from rice husk, exposing the OH functional group on the surface of the rice husk. Thus, sodium hydroxide treatment of CF could follow

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a reaction scheme as given by CF  OH þ NaOH ! CF  O Naþ þ H2 O

ð12Þ

Hence, the rate-limiting step of the Hg(II) sorption process depends on mercury diffusion in either the boundary film (external mass transfer) or pores within CFB, which could be evaluated by using the intraparticle diffusion model of Weber and Morris, given by qt ¼ kid t0:5

ð13Þ

where kid is the intraparticle diffusion rate constant (mg=g  min0.5). This model assumes that the external resistance to mass transfer (film diffusion) is either insignificant or significant but only for very short period at the beginning of diffusion, and, therefore, intraparticle diffusion is the rate-limiting step. The plot of q versus t0.5 will result in a straight line with a slope that equals kid and an intercept equal to zero. Generally, the plot of q versus t0.5 gave multi-straight lines, indicating that two or more steps occurred. The first sharp straight line indicates external surface sorption or boundary layer diffusion (film diffusion) and the second straight line indicates the gradual sorption stage, where intraparticle diffusion was a rate-controlled process. The third straight line indicates the final equilibrium stage where intraparticle diffusion started to slow down due to extremely low adsorbate concentrations in the solution (Kumar and Gaur, 2011; Ofomaja, 2010). Figure 9 shows the plot of q versus t0.5 for Hg(II) sorption on CFB kinetic data with the two straight lines that represent intraparticle diffusion and the final equilibrium stage. These straight lines do not pass through the origin and do not provide a suitable fitting to the kinetic data (Figure 8). This means that intraparticle diffusion was not a rate-controlling mechanism and thus the Hg(II) sorption process was governed by the external surface sorption and intraparticle diffusion. The intraparticle diffusion rate constants, kid, as obtained from the slopes of two straight lines, are listed in Table IV.

Figure 9. Intraparticle diffusion plots of Hg(II) sorption onto CFB. The experimental data are the same as in Figure 4.

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K. Johari et al. Table IV. Intraparticle rate parameters and diffusion coefficients for Hg(II) sorption onto CFB at 30  0.5 C CFB Kinetic models

CF-Pristine

Intraparticle diffusion model (Weber-Morris) Kid1 (mg=g  min0.5) 2.195 R2 0.990 Kid2 (mg=g  min0.5) 1.202 R2 0.996 Pore diffusion model (Boyd plot) D1 (m2=min)  1011 3.241 R2 0.928 Film diffusion model D2 (m2=min)  1012 6.450 R2 0.926

CF-NaOH 2.316 0.993 1.040 0.999 4.457 0.949 10.940 0.940

Generally, the intraparticle process is assumed to be controlled by pore and film diffusion within the biosorbent (Ofomaja, 2010). In order to calculate the pore diffusion (D1), the kinetic data was analyzed using the Boyd plot (Kumar and Gaur, ¨ nal et al., 2007): 2011; Ofomaja, 2007; O F values > 0:85 Bt ¼ 0:4997  lnð1  F Þ

ð14Þ

p p F values < 0:85 Bt ¼ ½ p  ðp  ðp2 F =3ÞÞ2

ð15Þ

where F ¼ qt=qe. The calculated Bt values were then plotted against time. The slope (B) of the linear Boyd plots was used to calculate the pore diffusion coefficient (D1) using Equation (16), assuming the biosorbent particle to be a sphere of radius a: B ¼ p2 ðD1 =a2 Þ

ð16Þ

The film diffusion coefficients (D2) were calculated from the slope of the plots of qt= qe versus t0.5, in which the data of the plots were calculated using: qt =qe ¼ 6ðD2 =pa2 Þ0:5  t0:5

ð17Þ

The pore diffusion (D1) and film diffusion (D2) coefficients are shown in Table IV. Figure 8 shows the Boyd plot and film diffusion fitting for the CFB kinetic data. It was noticed that the plots did not provide an accurate fitting to the CFB kinetic data as compared to the pseudo-second-order model. The calculated results indicated that the film diffusion coefficients for CF-Pristine and CF-NaOH (6.455  1012 and 10.950  1012 m2=min) were relatively lower than the pore diffusion coefficients (3.241  1011 and 4.457  1011 m2=min). This led to the suggestion that the Hg(II) transfer process occurred more slowly inside the boundary film than in the pores within CFB. Furthermore, it was believed that this could confirm the dominating nature of the film diffusion mechanism on the overall Hg(II) sorption process of CFB. Similar results were reported for the adsorption of Cd(II),

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Cu(II), and Pb(II) by various types of biosorbents (Garcia-Reyes and RangelMendez, 2010; Kumar and Gaur, 2011; Ofomaja, 2010). They reported that external mass transfer was the dominant rate-limiting step with calculated film diffusion coefficients in the range of 109 to 1011 m2=min. However, pore diffusion controlling the sorption process was also reported, such as in the sorption of Hg(II), As(II), and Pb(II) ions onto thiolated coconut fibers (Igwe et al., 2008). It is assumed that NaOH treatment of CF may lead to an increase of the intraparticle diffusion rate constant, kid, pore diffusion coefficient, D1, and film diffusion coefficient, D2, indicating a faster kinetic process. However, as previously discussed, the sorption capacity observed for CF-Pristine was significantly higher than for CF-NaOH. The film thickness for both CF-Pristine and CF-NaOH, which were 1.955  104 and 1.954  104 m, respectively, might have had insignificant effects on the diffusional sorption process. Mohan et al. (2008) and Okoye et al. (2010) also reported that the film thickness of the biosorbent was about 103 cm. Furthermore, the higher density of functional groups on the CF-NaOH surfaces, resulting from NaOH treatment, might lead to a larger interaction between Hg(II) and the surface functional groups, hence promoting faster reaction kinetics. Sorption Temperature Dependence The effect of temperature on Hg(II) sorption onto CFB was analyzed using temperatures of 30 , 45 , and 60 C. The results revealed that the sorption capacity decreased when the temperature increased from 30 to 60 C. These results indicated that Hg(II) sorption onto CFB is exothermic in nature. The Gibbs free energy of the sorption reaction, DGo (kJ=mol), is given by: DG o ¼ RT ln Ka

ð18Þ

where Ka is considered to be the sorption equilibrium constant, which is obtained by Ka ¼ Ce(sorbent)=Ce(solution). The Gibbs free energy change is also related to the change in enthalpy, DHo (kJ=mol), and entropy, DSo (kJ=mol  K), and at constant temperature, as given by: DG o ¼ DH o  TDSo

ð19Þ

Combining Equation (18) and (19) gives the van ‘t Hoff equation: ln Ka ¼ ðDG o =RTÞ ¼ ðDSo =RÞ  ðDH o =RTÞ

ð20Þ

where T is the absolute temperature (K) and R is the universal gas constant (8.314 J= mol  K). The values of DGo, DSo, and DHo were calculated for Hg(II) sorption on CFB, and the results are shown in Table V. The negative value of DGo indicated the spontaneous nature of Hg(II) sorption onto CFB. The values of DGo for CFPristine and CF-NaOH indicated that more negative values of DGo were observed for CF-Pristine, which implied a greater driving force of Hg(II) sorption than that for alkali-treated CF-NaOH. The positive value of DSo suggests that the randomness of the solid-solution interface increased during the Hg(II) sorption process, and the negative value of DHo indicated that the sorption reaction is an exothermic process. These results showed that the favorable behavior of the Hg(II) sorption process at all

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Table V. Thermodynamic parameters for Hg(II) sorption onto CFB Biosorbent

Temperature ( C)

CF-Pristine

30 45 60 30 45 60

CF-Pristine

DGo (kJ=mol)

DHo (kJ=mol)

DSo (kJ=mol  K)

7.2272 5.0568 1.1788 5.5725 1.3910 0.8528

68.0251

8.1168

54.8415

8.1577

temperatures agreed well with the literature reviewed by Doke and Khan (2013). Typically, the enthalphy for physical adsorption lies in the ranges from 4 to 40 kJ=mol, and from 40 to 800 kJ=mol for chemisorption (Crini and Badot, 2008). As shown in Table V, the DHo values of CF-Pristine and CF-NaOH were found to be 68.025 and 54.8415 kJ=mol, respectively, which could be attributed to a chemical sorption process. As previously discussed, similar results were obtained from the D-R and Temkin isotherm models’ analysis. Thus, the results confirmed that the Hg(II) sorption mechanism onto CFB followed the chemical sorption process.

Conclusion Here, we have succesfully shown that CF has potential to be employed as a Hg(II) biosorbent. Modified CF, which was prepared through NaOH treatment (CFNaOH), showed a lower Hg(II) sorption capacity than that of pristine CF (CFPristine). The Langmuir isotherm model was the best fitting model with a sorption capacity, qmax, of CF-Pristine and CF-NaOH of 166.67 and 142.86 mg=g, respectively. The kinetics of the Hg(II) sorption process onto CFB obeyed the pseudosecond-order kinetic model over the whole range of the sorption process. Based on the diffusion kinetic models consulted, it was found that intraparticle and external mass transfer processes affected the Hg(II) sorption process of CFB, in which film diffusion was the rate-limiting step. These results demonstrated the potential application of CF as a candidate for Hg(II) sorption. At present, further study is in progress that aims to explore the potential of developing CF for Hg(II) removal from process streams and mercury-contaminated production wastewater and effluents in the oil and gas industry.

Funding The financial support from the Ministry of Agriculture (MOA), Malaysia under the eScience Research Program (Project No. 05-01-06-SF1006), Research University Grant (GUP 00H63), and the 2009 ExxonMobil Research Grant is gratefully acknowledged.

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