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International Biodeterioration & Biodegradation 92 (2014) 12e19

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International Biodeterioration & Biodegradation

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Mercury removal from aqueous solution by dried biomass ofindigenous Vibrio parahaemolyticus PG02: Kinetic, equilibrium, andthermodynamic studies

Seyed Ali Jafari, Sama CheraghiDepartment of Biotechnology, Persian Gulf Research Institute, Persian Gulf University, P.O. Box: 75169, Bushehr, Iran

a r t i c l e i n f o

Article history:Received 1 June 2013Received in revised form17 January 2014Accepted 20 January 2014Available online

Keywords:BiosorptionVibrio parahaemolyticusIsothermMercuryIon-exchangeKinetic

E-mail address: sajafari.pgrsc@yahoo.com (S.A. Jaf

http://dx.doi.org/10.1016/j.ibiod.2014.01.0240964-8305/� 2014 Elsevier Ltd. All rights reserved.

a b s t r a c t

In this study, for the first time the potential use of dried Vibrio parahaemolyticus PG02 to remove mercuryfrom synthetic effluent was investigated by considering equilibrium, kinetic, and thermodynamic as-pects. The results indicated that Hg2þ biosorption was best described by the pseudo-second order model.In addition, it was found that intraparticle diffusion was not the sole rate-limiting step. The ion-exchangemechanism was a predominant biosorption mechanism in the first 15 min of contact. The Langmuirisotherm better described the equilibrium data of Hg2þ biosorption than the Freundlich isotherm. Ac-cording to this model, the maximum biosorption capacity was found to be 9.63 � 10�4 mol g�1 at op-timum conditions (pH ¼ 6.0 and temperature ¼35 �C).

� 2014 Elsevier Ltd. All rights reserved.

1. Introduction

Mercury is one of the most toxic heavy metal in the envi-ronment, one that easily accumulates in living tissues. In 2000close to 2190 tons of mercury were released into the environ-ment by anthropogenic activities (Sinha et al., 2012). Sinceelemental mercury has a long residence time of at least a fewmonths, as well as a high vapor pressure at room temperature, itcan evaporate and affect remote locations (Lindqvist and Rodhe,1985). Therefore, there is a need for novel, low-cost technologiesfor mercury removal in order to reduce its concentration in theenvironment.

The use of living or nonliving microorganisms as biosorbentshas emerged recently as an efficient, eco-friendly, and inexpen-sive alternative for removal of trace levels of heavy metals or oforganic contaminants from contaminated effluents (Esmaeiliet al., in press). Biosorption requires the binding of heavymetals or of other chemicals to the surface ligands of microor-ganisms (Rezaee et al., 2006). Due to the complex anddiverse structure of microorganism cell walls, the details of

ari).

biosorption mechanisms are still unknown (Mo and Lian, 2011).The use of dead cells for metal removal offers advantages overusing living cells since there is no need to provide nutrients forgrowth or maintenance in the solution to be treated, regenera-tion of the biosorbents may be obtained by simpler procedures,and contamination control is not required (Rezaee et al., 2006;Das et al., 2008; Plaza et al., 2011). Most data in the literatureindicate that dead biomass has a sorption capacity higher thanthat of live biomass (Kacar et al., 2002; Bayramo�glu and Arıca,2008; Velásquez and Dussan, 2009; Li et al., 2010; Huanget al., 2013). Biosorbent selection can directly affect the econ-omy of a biosorption process. A good biosorbent should be amicrobial strain that occurs naturally in the environment to betreated or in which the waste is to be disposed; it should not begenetically modified; and it should be easy to maintain and becapable of rapid growth in inexpensive media (Vieira andVolesky, 2000).

This study was the first to investigate dried Vibrio sp. cells asbiosorbents for removing Hg2þ from aqueous solutions underdifferent conditions of pH, metal concentration, and temperature.Pseudo-first- and pseudo-second-order rate equations, and Lang-muir and Freundlich isotherm models, as well as thermodynamicstudies, were performed to characterize Hg2þ biosorption by theorganism.

S.A. Jafari, S. Cheraghi / International Biodeterioration & Biodegradation 92 (2014) 12e19 13

2. Material and methods

2.1. Materials

All chemicals and mercuric chloride (HgCl2) were of analyticalgrade and purchased from Merck. Solutions were prepared withdouble-distilled water. All glassware was soaked in 10% nitric acidand rinsed several times with distilled water prior to the experi-ments to avoidmetal contamination. A stock solution (1000mg l�1)of Hg2þ was prepared by dissolving the required quantity of mer-cury chloride (HgCl2) in double distilled water. The pH of the so-lutions was adjusted using 1 M NaOH or 1 M HCl.

2.2. Preparation of biosorbent

The Vibrio parahaemolyticus PG02 (GenBank accession numberKC990033) strain that was used in this study as biosorbent was iso-latedpreviously fromBushehr (Iran)coastal sediments and identifiedas the most mercury-resistant microbial strain in that sample (Jafariet al., in press). Live cells of PG02 have a good potential to removemercury ions from aqueous solutions (Jafari et al., in press).

Cells of V. parahaemolyticus PG02 were inoculated into a 500-mlErlenmeyer flask containing 200ml tryptic soy broth (TSB) mediumcontaining, per liter of double distilled water, 17.0 g peptone fromcasein, 3.0 g peptone from soymeal, 2.5 g D(þ)glucose, 5.0 g sodiumchloride, 2.5 g di-potassium hydrogen phosphate, and 25 g of NaCl.The culture was incubated in a conventional shaker at 160 rpm for8 h at 35 �C. Initial pH was 7.2. The flask content was then trans-ferred to a 10-L bioreactor (Electrolab, ferMac 360) containing 4 L ofTSB culture medium and agitated (160 rpm) at 35 �C. The pH waskept constant at 7.0 during the incubation period. The biomass washarvested after 48 h, at the stationary phase of growth, by centri-fugation at 4000 rpm for 30 min at room temperature. The cellpellets were subsequently washed twice with double distilledwater and dried in a conventional oven at 60 �C for 10 h (Wanget al., 2010).

2.3. Batch biosorption

A series of batch experiments were carried out to determine theoptimum conditions for biosorption. These experiments wereperformed in 250-ml Erlenmeyer flasks containing 100 ml Hg2þ

solution. The pH of the solutions was initially adjusted to 6.0 (notcontrolled during the experiments) and the temperature was fixedat 35 �C. The effect of contact time on biosorption kinetics wasstudied with four different metal solutions (10, 40, 100, and200 mg l�1 Hg2þ). Samples were removed at regular time intervalsup to 120min. The effect of pHwas investigated in the range of 3e7in flasks containing 10 mg l�1 Hg2þ. Some controls including onlythe mercury solution without biosorbent were run simultaneouslywith the biosorption experiments in order to investigate the effectof glassware on mercury removal. In all the procedures, the solu-tions were in contact with 1.0 g l�1 of biosorbent and were shakenat 160 rpm. The biomass was harvested by centrifugation at5000 rpm for 30 min. The residual Hg2þ concentration in the su-pernatant was analyzed with a flameless atomic absorption spec-trophotometer (PG instruments, AA500). All the adsorptionexperiments were performed in duplicate and confidence intervalsof 95% were determined for each set of samples. The biomasssorption capacity, q (mol g�1), was calculated using Eq. (1) as fol-lows (Li et al., 2010):

q ¼�Ci � Cf

�$V

m(1)

where Ci and Cf are the initial and residual mercury concentrations(mol l�1), respectively, V is the solution volume (in liters), and m isthe dry weight of the biomass (in grams).

2.4. Kinetics and mechanisms of biosorption

The pseudo-first-order rate expression of Lagergren(Bayramo�glu and Arıca, 2008) and the pseudo-second-order rateequation have been used for modeling the adsorption of Hg2þ ionsin the range of 10e200 mg l�1 Hg2þ during the initial 60 min (El-Sikaily et al., 2007; Esmaeili et al., in press).

The contribution of the ion-exchange mechanism to biosorptionwas evaluated by recording the initial and final pH of the solutionsafter 60 min of contact, in addition to data evaluation by theequations proposed by the rate equation of Boyd et al. (1947). Thecontrol runs in these experiments were included in the mixture ofthe same concentration of biosorbent in water without any mer-curic ions at different initial pH values in order to assess the effectof biosorbent on the changes in the final pH of the medium.

The intraparticle diffusion mechanism was also investigated byevaluating the linearity of the sorption capacities over the squareroot of time, as will be described in Section 3.5.

2.5. Equilibrium isotherms

The biosorption isotherm experiments were performed asdescribed above under the optimum conditions determined inprevious experiments. The pH was kept constant during the ex-periments. A wide range of initial mercury concentrations (from 10to 350 mg l�1) was employed in these experiments. The Langmuirand Freundlich sorption isotherms were evaluated for analysis ofthe equilibrium data.

2.6. Thermodynamic studies

The adsorption equilibrium experiments were conducted attemperatures between 15 and 45 �C at pH 6.0. The thermodynamicparameters enthalpy change (DH0), entropy change (DS0), and freeenergy change of the sorption (DG0) were calculated for giventemperatures.

3. Results and discussion

3.1. Effect of contact time on Hg2þ biosorption

Contact time is one of the important parameters for successfulbiosorption. Experimental studies were carried out with varyinginitial mercury ion concentrations (10, 40, 100, and 200 mg l�1

Hg2þ) using 1.0 g l�1 adsorbent at 35 �C and an initial pH of 6. Asshown in Fig. 1, the equilibrium time of Hg2þ biosorption onto V.parahaemolyticus PG02 cells was independent of the initial Hg2þ

concentrations. For all mercury concentrations, the equilibriumwasreached after 60 min of contact. After that, mercury concentrationin the solution did not significantly change with time. At all con-centrations most of the metal removal occurred within the first15 min of contact. Thereafter the rate of removal decreased andfinally reached equilibrium. The initial fast uptake is probably dueto the large number of vacant binding sites on the solid surface, thehigh initial Hg2þ concentration in the solution, and consequentlythe high driving force between solution and solid surface. Theslower subsequent phase was due to the saturation of metalbinding sites that increased the repulsive forces between the Hg2þ

ions on the solid and liquid phases (Joo et al., 2010; Tunali Akaret al., 2012). Kacar et al. (2002) reported the biosorption equilib-rium time of 60 min for Hg2þ removal by immobilized inactive and

Fig. 2. Effect of initial pH on Hg2þ removal capacity for Vibrio parahaemolyticusPG02 cells. Biomass concentration 1.0 g l�1, initial mercury concentration 10 mg l�1,contact time 60 min, and temperature 35 �C. Confidence intervals ¼95%.

S.A. Jafari, S. Cheraghi / International Biodeterioration & Biodegradation 92 (2014) 12e1914

live fungi Phanerochaete chrysosporium. The equilibrium Hg2þ

sorption time for inactive Bacillus sp. biomass and heat-inactivatedpellets of Lentinus edodes was 60 min (Green-Ruiz, 2006;Bayramo�glu and Arıca, 2008). For subsequent experiments, thecontact time was standardized to 60 min.

The small loss of mercury in the controls (120 min of contacttime) of below 1% reflected the adsorption of mercury on the sur-face of the glassware. Stas’ et al. (2004) previously showed thatheavymetal ions could be adsorbed at different rates on the surfaceof glassware depending on the type of glass used. Jafari et al. (inpress) reported mercury depletion below 6% during 50 h of con-tact time between Hg2þ solution and the glass surface. Sinha andKhare (2012) also observed an almost 13% reduction in mercuryconcentration during 144 h of contact between mercury solution,without bacteria, and a glass container.

3.2. Effect of pH on Hg2þ biosorption

Since ion-exchange is one of the possible mechanisms for Hg2þ

biosorption, the uptake of mercury ions by biosorbent is expectedto be highly pH dependent. In general, the influence of pH onbiosorption is closely related to the ionic states of functional groupson the cell wall as well as of the metal in solution (Lacher andSmith, 2002b; Herrero et al., 2005). The experiments were per-formed at different pH values (ranging from 3.0 to 7.0) at 35 �C.

Fig. 2 shows that the initial pH of the solution had significanteffects on mercury uptake by dried PG02. The maximum Hg2þ

removal (4.44 � 10�5 mol g�1) at pH 6 was reduced to3.89 � 10�6 mol g�1 at pH 3. High sorption competition betweenHg2þ and protons at low pH led to a decrease in metal sorption ca-pacity (Green-Ruiz, 2006; Ofomaja and Ho, 2007; Bayramo�glu andArıca, 2008; Khoramzadeh et al., 2013). The specific surface ligandsthat contributed to sorption of Hgþ2 ions were, possibly, neutralizedat acidic pH (Lacher and Smith, 2002b). Several authors reportedadverse effects of protonation of cell wall components on the bio-sorption capacity of biomass (Kacar et al., 2002; Cain et al., 2008;Lesmana et al., 2009). On the other hand, as described above, themetal speciation in the solution severely affects themetal adsorptionprocess, especially for inorganic mercury. As Herrero et al. (2005)have demonstrated, mercury in solutions at pH values less than 6is mainly complexed with chloride while at higher pH values mer-curyehydroxide complexes predominate, and these are less prone toattraction by the biosorbent. However, the fraction of mercurycomplexed to chlorides in acidic medium directly depends on theconcentration of Cl� ions in themedium. The predominant species inthe presence of high concentrations of Cl� can be HgCl2, HgCl3- , and

Fig. 1. Effect of contact time on Hg2þ removal by dried Vibrio parahaemolyticus PG02 atinitial concentrations of 10 (A), 40 (-), 100 (:), and 200 (�) mg l�1 Hg2þ. Biomassconcentration 1.0 g l�1, pH 6.0, and temperature 35 �C.

small amounts of HgCl42�.The predominant mercury species is Hg2þ

in the absence or at low concentration of chloride ions in the me-dium. Miretzky et al. (2005) and Zhang et al. (2005) confirmed thisdistribution pattern in their studies on the speciation of mercury as afunction of pH in the presence of trace and high quantities of Cl�

ions. In the present study, complex formation is probably the reasonfor decrease in metal uptake at pH values higher than 6 (Green-Ruiz,2006; Esmaeili et al., in press). These complexes turned the mediumturbid. The control run, which included only metal solution (withoutbiosorbent), showed a similar turbidity at the same pH value.

Cain et al. (2008) and Plaza et al. (2011) reported an optimumpHvalue of 6 for mercury biosorption by active and inactive bacterial,fungal, and algal biosorbents. Optimum pH values in the range of4.5e6 were also reported for biosorption of Hg2þ (Kacar et al.,2002; Green-Ruiz, 2006; Bayramo�glu and Arıca, 2008).

An alternative explanation for the observed trend of Hg2þ

sorption is the alteration of the ion-exchange power by modifica-tion of the initial pH of the solution. For this purpose, the final pH ofthe solutions was also recorded after 60 min of contact and thechanges in the proton ion concentration (DHþ) were calculated bysubtracting the initial proton ion concentration, [Hþ], from the finalone.

Table 1 shows that the equilibrium sorption capacity value, qe,was proportional to the change in hydrogen ion concentration ofthe solution (DHþ). This suggests that higher DHþ improved theion-exchange potential, providing higher sorption capacity.Ofomaja and Ho (2007) also reported high sorption capacities athigh DHþ values. The pH of the solutionwas reduced because of thereplacement of Hþ ions on the biomass surface by Hg2þ ions fromthe solution. However, the final pH of the control runs includingonly the bacterial biosorbents (without mercury) did not signifi-cantly change. If there was any proton replacement in the solution,in the absence of mercury ions, presumably the Hþ ions of func-tional groups were replaced by Hþ ions from the bulk.

Eq. (2) represents the linear relationship between equilibriumcapacity and DHþ. The ion-exchange rate was15.04 mg gmM�1 with a high coefficient of correlation (0.978).

Table 1Changes in hydrogen ion concentration at different initial pH of the solution. Atinitial mercury concentration of 10 mg l�1 and 60 min of contact.

Initial pH Final pH DHþ (mM) qe(mol g�1)

3 3.2 �0.3690 3.89 � 10�6

4 3.8 0.0585 3.32 � 10�5

5 3.9 0.1159 3.89 � 10�5

6 3.9 0.1249 4.44 � 10�5

7 4 0.0999 3.78 � 10�5

S.A. Jafari, S. Cheraghi / International Biodeterioration & Biodegradation 92 (2014) 12e19 15

qe ¼ 15:04DHþ þ 6:253 (2)

This value is comparable with similar cases reported in theliterature: 33.22 mg gmM�1 for biosorption of lead(II) onto man-sonia wood sawdust (Ofomaja, 2010), 0.289 mg gmM�1 for cad-mium(II) adsorption onto coconut copra meal (Ofomaja and Ho,2007), and 10.60 mg gmM�1 for lead(II) transfer onto tree fern(Ho, 2005).

3.3. Ion-exchange model

The contribution of the ion-exchange mechanism to the ab-sorption of mercury ions onto the dried biomass of V. para-haemolyticus PG02 can be assessed from the variation of initialmercury concentration (10e200 mg l�1). Boyd et al. (1947) devel-oped a rate equation, that considered rates of ion-exchangeadsorption by organic zeolites as follows:

logð1� FÞ ¼ ��

S2:303

�t (3)

where F is the fractional attainment of equilibrium, F¼qt/qe, and S(min�1) is a constant. The application of this model was tested byOfomaja and Ho (2007) and Ho (2005) for biosorption of lead(II).This model was used in this study for investigating the involvementof ion-exchange reactions in the biosorption of Hg2þ onto driedbiomass of PG02. Fig. 3 shows a good fit of the experimental data bytheoretical curves in the first 15 min of contact at all metal con-centrations. After that, the experimental data do not follow themodel predictions. Boyd et al. (1947) rate equation constants (S) atall initial mercury concentrations for this period are presented inTable 2, along with their standard error values. The decrease of therate constant, S, at increasing metal concentration, is in accordancewith the findings of Ofomaja (2010).

Therefore, according to the Boyd et al. (1947) model, ion-exchange mechanisms drove biosorption of mercury to driedbiomass of V. parahaemolyticus PG02 in the first 15 min of contact.This behavior has been observed in other metal sorption studies(Ho, 2005; Ofomaja and Ho, 2007; Ofomaja, 2010).

3.4. Biosorption rate kinetics and the effect of initial Hg2þ

concentration

The effect of increasing initial Hg2þ concentration on the ki-netics of mercury biosorptionwas investigated by the use of several

Fig. 3. Ion-exchange biosorption kinetics of Hg2þ onto dried Vibrio parahaemolyticusPG02 at various initial Hg2þ concentrations. Biosorbent concentration 1.0 g l�1, pH 6.0,and temperature 35 �C during 20 min of contact.

initial mercury concentrations (10, 40, 100, and 200mg l�1 Hg2þ) atan initial pH of 6 and temperature of 35 �C for 60 min of contact.The first-order rate equation of Lagergren is one of the most widelyused relationships for modeling the sorption of a solute from aliquid solution. The linear form of Lagergren’s pseudo-first-orderrate equation is expressed as follows (Bayramo�glu and Arıca, 2008):

logðqe � qtÞ ¼ log qe � k12:303

t (4)

where qt and qe are the amounts of Hg2þ adsorbed at time t (inminutes) and equilibrium (mol g�1), respectively, and k1 is the rateconstant of the pseudo-first-order adsorption process (min�1). Thevalues of k1, and qe are determined from the slope and intercept ofthe plot of log (qe�qt) against t, respectively.

The linear form of the pseudo-second-order kinetic model isgiven as Eq. (5) (Tunali Akar et al., 2012):

tqt

¼ 1k2q2e

þ 1qe

t (5)

where k2 is the equilibrium rate constant for pseudo-second-orderbiosorption (g mol�1 min�1). Values of k2 and qe were calculatedfrom the plot of t/qt against t.

Fig. 4a and b represent the pseudo-first- and second-order plots,respectively, for all initial mercury concentrations. It is clear thatthe pseudo-second-order model describes the experimental datamuch better for the entire adsorption process (Fig. 4b).Khoramzadeh et al. (2013), however, reported that the first-orderequation of Lagergren is applicable only over the initial 20e30 min of the sorption process and not for the entire range ofcontact time investigated in this study.

The kinetic parameters for both models along with their co-efficients of correlations (R2) were tabulated in Table 2. The pseudo-second-order rate constant (k2) decrease from 8150.4 to407.1 g mol�1 min�1 with an increase in the initial metal concen-tration from 10 to 200 mg l�1 is consistent with the findings ofother researchers (El-Sikaily et al., 2007; Ofomaja, 2010). The co-efficient of correlation (R2) of the pseudo-second-order model (Eq.(5)) for all the studied concentrations (all above 0.992) was higherthan that of the pseudo-first order model (Eq. (4)). In addition,pseudo-first-order theoretical qe values were not appropriate fordescription of the adsorption behavior. This suggests that Hg2þ

biosorption by V. parahaemolyticus PG02 was not a first-order re-action. According to Sinha et al. (2012), if the experimental qe is notequal to the theoretical qe, the considered kinetic equation will notcorrespond to experimental data, even if the plot has a high coef-ficient of correlation.

Analysis of the experimental data by the pseudo-second-ordermodel suggests that the rate-limiting step was chemisorption(Lesmana et al., 2009). If the relationship between initial metalconcentration and rate of adsorption is not linear (as seen inTable 2) then other factors such as intraparticle diffusion may limitthe adsorption process (El-Sikaily et al., 2007).

3.5. Intraparticle diffusion

Since the Eqs. (4) and (5) cannot replicate the diffusion mech-anism satisfactorily, the intraparticle diffusion model was alsoevaluated (Gialamouidis et al., 2010):

qt ¼ kintt0:5 (6)

where kint is the intraparticle diffusion rate constant(mol g�1 min�0.5) and is the slope of the plot of qt versus t0.5

Table 2Pseudo-first order, Pseudo-second order, intraparticle diffusion, and ion-exchange model parameters for Hg2þ biosorption onto dried biomass of Vibrio parahaemolyticus PG2at different initial concentrations.

Ci (mg l�1) qe,exp(mol g�1)

Pseudo-first order model Pseudo-second order model Intraparticle diffusiona Ion-exchange modela,b

k1 (min�1) qe,model

(mol g�1)R2 k2 (g mol�1 min�1) qe,model

(mol g�1)R2 kint (mol g�1 min�0.5) R2 S (min�1) R2

10 4.53 � 10�5 0.061 2.4 � 10�5 0.917 8150.4 4.6 � 10�5 0.996 3.22 � 10�6 � 7.5 � 10�7 0.785 0.1252 � 0.002 0.99840 1.76 � 10�4 0.049 8.6 � 10�5 0.834 2165.8 1.8 � 10�4 0.995 1.25 � 10�5 � 3.6 � 10�6 0.702 0.1176 � 0.000 0.999100 4.18 � 10�4 0.053 2.2 � 10�4 0.839 755.2 4.2 � 10�4 0.993 3.38 � 10�5 � 9 � 10�6 0.738 0.1157 � 0.002 0.998200 7.34 � 10�4 0.047 4.2 � 10�4 0.903 407.1 7.3 � 10�4 0.994 5.59 � 10�5 � 1.2 � 10�5 0.814 0.1078 � 0.003 0.996

a These parameters were provided � standard error values (Uncertainty).b For the first 15 min of contact.

S.A. Jafari, S. Cheraghi / International Biodeterioration & Biodegradation 92 (2014) 12e1916

(Fig. 5). The values of kint, at different metal concentrations areshown in Table 2 along with their standard error values. The re-sults indicate that the intraparticle diffusion rate increased from3.22 � 10�6 to 5.59 � 10�5 mol g�1 min�0.5 with increase of theinitial mercury concentration from 10 to 200 mg l�1. Ofomaja(2010) claimed that as initial metal concentration is increased,intraparticle diffusion becomes more important, because a higherinitial metal concentration produces a stronger driving force fordiffusion.

According to Fig. 5, the overall line for each metal concentrationdoes not pass through the origin. El-Sikaily et al. (2007) andOfomaja (2010) reported that if the qt versus t0.5 plots do not passthrough the origin, then the intraparticle diffusion is not the solerate-limiting step. The poor R2 values for all initial mercury con-centrations for intraparticle diffusion (Table 2) indicate that theprocess of mercury biosorption on dried biomass of V. para-haemolyticus PG02 occurred by surface biosorption and intra-particle diffusion.

Fig. 4. Plots of the (a) pseudo-first-order and (b) pseudo-second-order equations for thebiosorption kinetics of Hg2þ onto dried Vibrio parahaemolyticus PG02 at various initialHg2þ concentrations. Biosorbent concentration 1.0 g l�1, pH 6.0, and temperature 35 �C.

This multi-linearity in the shape of the intraparticle diffusionplots has also been observed by Ofomaja (2010) in the biosorptionof lead(II) onto mansonia wood sawdust and by Gialamouidis et al.(2010) for biosorption of Mn(II) by Pseudomonas sp., Staphylococcusxylosus, and Blakeslea trispora cells.

3.6. Biosorption isotherm

In order to obtain information about the equilibrium, themaximum sorption capacity, and the affinity of the biosorbent tothe specific metal ions, and also for comparing the capacity ofdifferent biosorbents with each other, Langmuir and Freundlichequilibrium isotherms have been traditionally used to evaluatesorption data. Equilibrium experiments were performed underpredetermined conditions with initial mercury concentrations inthe range of 10e350 mg l�1 and 60 min of contact. The Langmuirmodel is based on a monolayer sorption on a homogeneous surfacewithout interaction between adsorbed molecules (Plaza et al.,2011). Eq. (7) shows the linear form of the Langmuir isotherm(Sinha et al., 2012):

Ceqe

¼ 1qmaxb

þ Ceqmax

(7)

where Ce is the equilibrium concentration (mol l�1), qmax is themaximum metal uptake (mol g�1) to form a complete monolayeron the surface and b is the Langmuir equilibrium constant(l mol�1) that represents the affinity between the sorbent andadsorbate. The qmax and b values can be determined by plottingCe/qe versus Ce. It should be emphasized that, in a strict theo-retical sense, Ce in Eq. (7) must be expressed as the molar con-centration (Liu, 2006).

Fig. 5. Intraparticle diffusion mechanism for Hg2þ biosorption by dried Vibrio para-haemolyticus PG02 at various initial Hg2þ concentrations. Biosorbent concentration1.0 g l�1, pH 6.0, and temperature 35 �C.

Fig. 6. (a) Langmuir and (b) Freundlich isotherm plots for Hg2þ biosorption onto driedVibrio parahaemolyticus PG02 cells at temperatures of 15 (A), 25 (-), 35 (:), and45 �C (�). Biomass concentration 1.0 g l�1, pH 6.0, and 60 min of contact.

Table 4Comparison of the maximum mercury biosorption capacity for different sorbents.

Sorbent Initial pH qmax (mol g�1) Reference

Bacillus sp. 4.5e6.0 3.94 � 10�5 Green-Ruiz (2006)Vibrio parahaemolyticus

PG026 9.63 � 10�4 This study

Chlamydomonasreinhardtii

6 6.1 � 10�4 Tüzün et al. (2005)

Ulva lactuca 5.5 4.22 � 10�4 Zeroual et al. (2003)Bacillus subtilis 5 3.56 � 10�4 Wang et al. (2010)Sugarcane Bagasse 4 1.78 � 10�4 Khoramzadeh

et al. (2013)Potamogeton natans Natural 1.13 � 10�3 Lacher and Smith

(2002a)Spirulina platensis 6 2.13 � 10�3 Cain et al. (2008)Walnut shell

activated carbons5 7.55 � 10�4 Zabihi et al. (2010)

Heat inactivatedPhanerochaetechrysosporium

5 8.58 � 10�4 Kacar et al. (2002)

S.A. Jafari, S. Cheraghi / International Biodeterioration & Biodegradation 92 (2014) 12e19 17

The Freundlich isotherm is an empirical model that proposes amonolayer sorption with heterogeneous energetic distribution ofactive sites (Plaza et al., 2011). The linear form of it is given by Eq.(8) (Sinha et al., 2012):

log qe ¼ log kF þ1nlog Ce (8)

where kF (mol g�1) and n are the Freundlich constant andFreundlich exponent, respectively (Gialamouidis et al., 2010). The kFvalue is an indicator of the biosorption capacity and 1/n representsthe surface heterogeneity and are calculated from the intercept andslope of a plot of log qe versus log Ce, respectively (Ertugay andBayhan, 2010; Tunali Akar et al., 2012).

Fig. 6 (a and b) illustrates the experimental equilibrium con-centrations for mercury removal by dried biomass of V. para-haemolyticus PG02 at different temperatures, as predicted by theLangmuir and Freundlich models, respectively. It is obvious that theLangmuir isotherm model describes the experimental data betterthan the Freundlich model does.

Table 3Langmuir and Freundlich constants along with thermodynamic parameters at differenthaemolyticus PG2.

Temperature Langmuir constants Freundlich consta

qmax (mol g�1) b (l mol�1) R2 kF (mol g�1)

15 �C 5.08 � 10�4 3954.28 0.989 0.03125 �C 8.20 � 10�4 4875.78 0.961 0.08335 �C 9.63 � 10�4 10816.32 0.998 0.06145 �C 8.86 � 10�4 8244.73 0.989 0.066

a These parameters were provided � standard error values (Uncertainty).

The maximum sorption capacity, qmax, at pH 6 and a tempera-ture of 35 �C was 9.63 � 10�4 mol Hg2þ g�1 biomass, which is highvalue in comparisonwith other reported sorbents for Hg2þ removal(Table 4).

In the case of the Freundlich model, usually for a goodadsorbent the value of n should be between 1 and 10. Largervalues of n indicate a strong bond between biosorbent and heavymetal (Li et al., 2010). The values achieved for nwere in the rangeof 1.53e1.78 for temperatures of 15e45 �C. Although it seemsthat these values were not large, they were a little higher thanthose reported in the literature as 0.55 (Khoramzadeh et al.,2013), 1.04 (Zhang et al., 2005), 1.10 (Esmaeili et al., in press),and 1.51 (Green-Ruiz, 2006) for Hg2þ biosorption by differentbiosorbents.

According to Table 3 the increase in temperature up to 35 �C ledto an increase in the sorption capacity from 5.08 � 10�4 to9.63 � 10�4 mol g�1, but higher temperatures reduced it to8.86 � 10�4 mol g�1. Presumably, increasing the temperatureslightly increased the migration of Hg2þ ions into the inner struc-ture of the biosorbent (Ofomaja, 2010). The significant decrease inbiosorption capacity above 35 �C was caused possibly by thedestruction of surface binding ligands of PG02 (Gialamouidis et al.,2010). These results are in accordance with results reported byGreen-Ruiz (2006).

3.7. Thermodynamic studies and the effect of temperature on Hg2þ

biosorption

In order to perform the thermodynamic studies, equilibriumexperiments were repeated at four different temperaturesd15, 25,35, and 45 �Cdat a pH of 6. The free energy change of theadsorption, DG0 kJ mol�1, as the fundamental criterion of sponta-neity, was determined from Eq. (9) (Gialamouidis et al., 2010):

temperatures for the biosorption of Hg2þ ions onto dried biomass of Vibrio para-

nts Thermodynamic parametersa

n R2 DG0 (kJ mol�1) DH0 (kJ mol�1) DS0 (j mol�1 K�1)

1.63 0.914 �29.45 18.97 � 1.44 167.93 � 4.81.53 0.879 �30.991.78 0.950 �34.071.68 0.902 �34.46

S.A. Jafari, S. Cheraghi / International Biodeterioration & Biodegradation 92 (2014) 12e1918

DG0 ¼ �RT ln K (9)

where R is the gas constant (8.314 J mol�1 K�1), T is the absolutetemperature, and K is the thermodynamic equilibrium constantwithout unit. The use of the Langmuir equilibrium constant, b,instead of the dimensionless thermodynamic equilibrium constantin Eq. (9) is incorrect, as reported byMilonji�c (2007) and Liu (2009).The b constant with the units l mol�1, can then K be madedimensionless by multiplying with 55.5, number of moles of waterper liter of solution (Milonji�c, 2007). The Langmuir equilibriumconstant with units of l mol�1 can be used directly for determina-tion of DG0, without any modification, only if neutral or weakcharge adsorbates or a dilute solution of charged adsorbate areused (Liu, 2009).

The free energy change of the adsorption is:

DG0 ¼ DH0 � TDS0 (10)

where DH0 is the enthalpy change (kJ mol�1) and DS0 is theentropy change (j mol�1 K�1), that were evaluated from theintercept and the slope of a plot of DG0 versus T, respectively(Table 3). The positive value of DH0 (18.97 kJ mol�1) indicatesthat the sorption of Hg2þ ions by dried biomass of PG02 wasendothermic (Lesmana et al., 2009; Liu, 2009). The increase inthe values of qmax with temperature (Table 3) confirms theendothermic nature of the process. The positive value forenthalpy change also suggests a high probability that adsorptionoccurred mainly by chemisorption. However, heats of chemi-sorption are generally in the range of 80e200 kJ mol�1, whereasthe enthalpy change in this work was in the range of heats ofcondensation, from 2.1 to 20.9 kJ mol�1. Therefore, the enthalpychange measured in this work suggests mercury adsorption ontodried biomass of PG02 occurred via physical adsorption (Liu,2009; Gialamouidis et al., 2010).

The free energy change values,DG0, were negative for all studiedtemperatures. This indicates that the biosorption process wasspontaneous (Lesmana et al., 2009). The positive value of DS0

(167.93 j mol�1 K�1) indicates an increased randomness at thesolid/solution interface during the adsorption of mercury ontobiomass and some structural changes in the cell wall.

4. Conclusions

The optimum conditions for mercury removal by V. para-haemolyticus PG02 were a pH 6 at 35 �C. The ion-exchange mech-anismwas a predominant mechanism in the first 15 min of contact.Only the pseudo-second-order model described the experimentaldata for the entire range of contact time and the intraparticlediffusion was not the sole rate-limiting step. According to theLangmuir model, the qmax value was calculated to be9.63 � 10�4 mol g�1 at optimum conditions (pH value of 6 andtemperature of 35 �C). Thermodynamic parameters confirmed thatthe physical adsorption process was endothermic and spontaneous,with a high positive value for DS0. The results suggest the PG02strain is a good adsorbent for mercury removal from aqueoussolutions.

Acknowledgment

This research was done with the financial support of theDepartment of Research Affairs of the Persian Gulf University ofBushehr, Iran, and the laboratory facilities of the Persian GulfResearch Institute (1146/1/1393-10/PGU/FC).

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