Mercury removal from aqueous solution by dried biomass of indigenous Vibrio parahaemolyticus PG02: Kinetic, equilibrium, and thermodynamic studies

International Biodeterioration & Biodegradation 92 (2014) 12e19

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Mercury removal from aqueous solution by dried biomass of indigenous Vibrio parahaemolyticus PG02: Kinetic, equilibrium, and thermodynamic studies Seyed Ali Jafari, Sama Cheraghi Department 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

a b s t r a c t

Article history: Received 1 June 2013 Received in revised form 17 January 2014 Accepted 20 January 2014 Available online

In this study, for the first time the potential use of dried Vibrio parahaemolyticus PG02 to remove mercury from synthetic effluent was investigated by considering equilibrium, kinetic, and thermodynamic aspects. 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-exchange mechanism was a predominant biosorption mechanism in the first 15 min of contact. The Langmuir isotherm better described the equilibrium data of Hg2þ biosorption than the Freundlich isotherm. According to this model, the maximum biosorption capacity was found to be 9.63  104 mol g1 at optimum conditions (pH ¼ 6.0 and temperature ¼35  C). Ó 2014 Elsevier Ltd. All rights reserved.

Keywords: Biosorption Vibrio parahaemolyticus Isotherm Mercury Ion-exchange Kinetic

1. Introduction Mercury is one of the most toxic heavy metal in the environment, one that easily accumulates in living tissues. In 2000 close to 2190 tons of mercury were released into the environment by anthropogenic activities (Sinha et al., 2012). Since elemental mercury has a long residence time of at least a few months, as well as a high vapor pressure at room temperature, it can evaporate and affect remote locations (Lindqvist and Rodhe, 1985). Therefore, there is a need for novel, low-cost technologies for mercury removal in order to reduce its concentration in the environment. The use of living or nonliving microorganisms as biosorbents has emerged recently as an efficient, eco-friendly, and inexpensive alternative for removal of trace levels of heavy metals or of organic contaminants from contaminated effluents (Esmaeili et al., in press). Biosorption requires the binding of heavy metals or of other chemicals to the surface ligands of microorganisms (Rezaee et al., 2006). Due to the complex and diverse structure of microorganism cell walls, the details of

E-mail address: [email protected] (S.A. Jafari). http://dx.doi.org/10.1016/j.ibiod.2014.01.024 0964-8305/Ó 2014 Elsevier Ltd. All rights reserved.

biosorption mechanisms are still unknown (Mo and Lian, 2011). The use of dead cells for metal removal offers advantages over using living cells since there is no need to provide nutrients for growth or maintenance in the solution to be treated, regeneration 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 literature indicate that dead biomass has a sorption capacity higher than
lu and Arıca, that of live biomass (Kacar et al., 2002; Bayramog 2008; Velásquez and Dussan, 2009; Li et al., 2010; Huang et al., 2013). Biosorbent selection can directly affect the economy of a biosorption process. A good biosorbent should be a microbial strain that occurs naturally in the environment to be treated or in which the waste is to be disposed; it should not be genetically modified; and it should be easy to maintain and be capable of rapid growth in inexpensive media (Vieira and Volesky, 2000). This study was the first to investigate dried Vibrio sp. cells as biosorbents for removing Hg2þ from aqueous solutions under different conditions of pH, metal concentration, and temperature. Pseudo-first- and pseudo-second-order rate equations, and Langmuir and Freundlich isotherm models, as well as thermodynamic studies, were performed to characterize Hg2þ biosorption by the organism.

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2. Material and methods

where Ci and Cf are the initial and residual mercury concentrations (mol l1), respectively, V is the solution volume (in liters), and m is the dry weight of the biomass (in grams).

2.1. Materials All chemicals and mercuric chloride (HgCl2) were of analytical grade and purchased from Merck. Solutions were prepared with double-distilled water. All glassware was soaked in 10% nitric acid and rinsed several times with distilled water prior to the experiments to avoid metal contamination. A stock solution (1000 mg l1) of Hg2þ was prepared by dissolving the required quantity of mercury chloride (HgCl2) in double distilled water. The pH of the solutions was adjusted using 1 M NaOH or 1 M HCl. 2.2. Preparation of biosorbent The Vibrio parahaemolyticus PG02 (GenBank accession number KC990033) strain that was used in this study as biosorbent was isolated previously from Bushehr (Iran) coastal sediments and identified as the most mercury-resistant microbial strain in that sample (Jafari et al., in press). Live cells of PG02 have a good potential to remove mercury ions from aqueous solutions (Jafari et al., in press). Cells of V. parahaemolyticus PG02 were inoculated into a 500-ml Erlenmeyer flask containing 200 ml tryptic soy broth (TSB) medium containing, per liter of double distilled water, 17.0 g peptone from casein, 3.0 g peptone from soymeal, 2.5 g D(þ)glucose, 5.0 g sodium chloride, 2.5 g di-potassium hydrogen phosphate, and 25 g of NaCl. The culture was incubated in a conventional shaker at 160 rpm for 8 h at 35  C. Initial pH was 7.2. The flask content was then transferred to a 10-L bioreactor (Electrolab, ferMac 360) containing 4 L of TSB culture medium and agitated (160 rpm) at 35  C. The pH was kept constant at 7.0 during the incubation period. The biomass was harvested after 48 h, at the stationary phase of growth, by centrifugation at 4000 rpm for 30 min at room temperature. The cell pellets were subsequently washed twice with double distilled water and dried in a conventional oven at 60  C for 10 h (Wang et al., 2010). 2.3. Batch biosorption A series of batch experiments were carried out to determine the optimum conditions for biosorption. These experiments were performed in 250-ml Erlenmeyer flasks containing 100 ml Hg2þ solution. The pH of the solutions was initially adjusted to 6.0 (not controlled during the experiments) and the temperature was fixed at 35  C. The effect of contact time on biosorption kinetics was studied with four different metal solutions (10, 40, 100, and 200 mg l1 Hg2þ). Samples were removed at regular time intervals up to 120 min. The effect of pH was investigated in the range of 3e7 in flasks containing 10 mg l1 Hg2þ. Some controls including only the mercury solution without biosorbent were run simultaneously with the biosorption experiments in order to investigate the effect of glassware on mercury removal. In all the procedures, the solutions were in contact with 1.0 g l1 of biosorbent and were shaken at 160 rpm. The biomass was harvested by centrifugation at 5000 rpm for 30 min. The residual Hg2þ concentration in the supernatant was analyzed with a flameless atomic absorption spectrophotometer (PG instruments, AA500). All the adsorption experiments were performed in duplicate and confidence intervals of 95% were determined for each set of samples. The biomass sorption capacity, q (mol g1), was calculated using Eq. (1) as follows (Li et al., 2010):

 q ¼

 Ci  Cf $V m

13

(1)

2.4. Kinetics and mechanisms of biosorption The pseudo-first-order rate expression of Lagergren
lu and Arıca, 2008) and the pseudo-second-order rate (Bayramog equation have been used for modeling the adsorption of Hg2þ ions in the range of 10e200 mg l1 Hg2þ during the initial 60 min (ElSikaily et al., 2007; Esmaeili et al., in press). The contribution of the ion-exchange mechanism to biosorption was evaluated by recording the initial and final pH of the solutions after 60 min of contact, in addition to data evaluation by the equations proposed by the rate equation of Boyd et al. (1947). The control runs in these experiments were included in the mixture of the same concentration of biosorbent in water without any mercuric ions at different initial pH values in order to assess the effect of biosorbent on the changes in the final pH of the medium. The intraparticle diffusion mechanism was also investigated by evaluating the linearity of the sorption capacities over the square root of time, as will be described in Section 3.5. 2.5. Equilibrium isotherms The biosorption isotherm experiments were performed as described above under the optimum conditions determined in previous experiments. The pH was kept constant during the experiments. A wide range of initial mercury concentrations (from 10 to 350 mg l1) was employed in these experiments. The Langmuir and Freundlich sorption isotherms were evaluated for analysis of the equilibrium data. 2.6. Thermodynamic studies The adsorption equilibrium experiments were conducted at temperatures between 15 and 45  C at pH 6.0. The thermodynamic parameters enthalpy change (DH0), entropy change (DS0), and free energy change of the sorption (DG0) were calculated for given temperatures. 3. Results and discussion 3.1. Effect of contact time on Hg2þ biosorption Contact time is one of the important parameters for successful biosorption. Experimental studies were carried out with varying initial mercury ion concentrations (10, 40, 100, and 200 mg l1 Hg2þ) using 1.0 g l1 adsorbent at 35  C and an initial pH of 6. As shown 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 equilibrium was reached after 60 min of contact. After that, mercury concentration in the solution did not significantly change with time. At all concentrations most of the metal removal occurred within the first 15 min of contact. Thereafter the rate of removal decreased and finally reached equilibrium. The initial fast uptake is probably due to the large number of vacant binding sites on the solid surface, the high initial Hg2þ concentration in the solution, and consequently the high driving force between solution and solid surface. The slower subsequent phase was due to the saturation of metal binding sites that increased the repulsive forces between the Hg2þ ions on the solid and liquid phases (Joo et al., 2010; Tunali Akar et al., 2012). Kacar et al. (2002) reported the biosorption equilibrium time of 60 min for Hg2þ removal by immobilized inactive and

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live fungi Phanerochaete chrysosporium. The equilibrium Hg2þ sorption time for inactive Bacillus sp. biomass and heat-inactivated pellets of Lentinus edodes was 60 min (Green-Ruiz, 2006;
lu and Arıca, 2008). For subsequent experiments, the Bayramog contact time was standardized to 60 min. The small loss of mercury in the controls (120 min of contact time) of below 1% reflected the adsorption of mercury on the surface of the glassware. Stas’ et al. (2004) previously showed that heavy metal ions could be adsorbed at different rates on the surface of glassware depending on the type of glass used. Jafari et al. (in press) reported mercury depletion below 6% during 50 h of contact time between Hg2þ solution and the glass surface. Sinha and Khare (2012) also observed an almost 13% reduction in mercury concentration 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 expected to be highly pH dependent. In general, the influence of pH on biosorption is closely related to the ionic states of functional groups on the cell wall as well as of the metal in solution (Lacher and Smith, 2002b; Herrero et al., 2005). The experiments were performed 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 significant effects on mercury uptake by dried PG02. The maximum Hg2þ removal (4.44  105 mol g1) at pH 6 was reduced to 3.89  106 mol g1 at pH 3. High sorption competition between Hg2þ and protons at low pH led to a decrease in metal sorption ca
lu and pacity (Green-Ruiz, 2006; Ofomaja and Ho, 2007; Bayramog Arıca, 2008; Khoramzadeh et al., 2013). The specific surface ligands that contributed to sorption of Hgþ2 ions were, possibly, neutralized at acidic pH (Lacher and Smith, 2002b). Several authors reported adverse effects of protonation of cell wall components on the biosorption capacity of biomass (Kacar et al., 2002; Cain et al., 2008; Lesmana et al., 2009). On the other hand, as described above, the metal speciation in the solution severely affects the metal adsorption process, especially for inorganic mercury. As Herrero et al. (2005) have demonstrated, mercury in solutions at pH values less than 6 is mainly complexed with chloride while at higher pH values mercuryehydroxide complexes predominate, and these are less prone to attraction by the biosorbent. However, the fraction of mercury complexed to chlorides in acidic medium directly depends on the concentration of Cl ions in the medium. The predominant species in the presence of high concentrations of Cl can be HgCl2, HgCl-3, and

Fig. 2. Effect of initial pH on Hg2þ removal capacity for Vibrio parahaemolyticus PG02 cells. Biomass concentration 1.0 g l1, initial mercury concentration 10 mg l1, contact time 60 min, and temperature 35  C. Confidence intervals ¼95%. 2þ small amounts of HgCl2 4 .The predominant mercury species is Hg in the absence or at low concentration of chloride ions in the medium. Miretzky et al. (2005) and Zhang et al. (2005) confirmed this distribution pattern in their studies on the speciation of mercury as a function of pH in the presence of trace and high quantities of Cl ions. In the present study, complex formation is probably the reason for decrease in metal uptake at pH values higher than 6 (Green-Ruiz, 2006; Esmaeili et al., in press). These complexes turned the medium turbid. The control run, which included only metal solution (without biosorbent), showed a similar turbidity at the same pH value. Cain et al. (2008) and Plaza et al. (2011) reported an optimum pH value of 6 for mercury biosorption by active and inactive bacterial, fungal, and algal biosorbents. Optimum pH values in the range of 4.5e6 were also reported for biosorption of Hg2þ (Kacar et al.,
lu and Arıca, 2008). 2002; Green-Ruiz, 2006; Bayramog An alternative explanation for the observed trend of Hg2þ sorption is the alteration of the ion-exchange power by modification of the initial pH of the solution. For this purpose, the final pH of the solutions was also recorded after 60 min of contact and the changes in the proton ion concentration (DHþ) were calculated by subtracting the initial proton ion concentration, [Hþ], from the final one. Table 1 shows that the equilibrium sorption capacity value, qe, was proportional to the change in hydrogen ion concentration of the solution (DHþ). This suggests that higher DHþ improved the ion-exchange potential, providing higher sorption capacity. Ofomaja and Ho (2007) also reported high sorption capacities at high DHþ values. The pH of the solution was reduced because of the replacement of Hþ ions on the biomass surface by Hg2þ ions from the solution. However, the final pH of the control runs including only the bacterial biosorbents (without mercury) did not significantly change. If there was any proton replacement in the solution, in the absence of mercury ions, presumably the Hþ ions of functional groups were replaced by Hþ ions from the bulk. Eq. (2) represents the linear relationship between equilibrium capacity and DHþ. The ion-exchange rate was 1 15.04 mg gmM with a high coefficient of correlation (0.978).

Table 1 Changes in hydrogen ion concentration at different initial pH of the solution. At initial mercury concentration of 10 mg l1 and 60 min of contact.

Fig. 1. Effect of contact time on Hg2þ removal by dried Vibrio parahaemolyticus PG02 at initial concentrations of 10 (A), 40 (-), 100 (:), and 200 () mg l1 Hg2þ. Biomass concentration 1.0 g l1, pH 6.0, and temperature 35  C.

Initial pH

Final pH

DHþ (mM)

qe(mol g1)

3 4 5 6 7

3.2 3.8 3.9 3.9 4

0.3690 0.0585 0.1159 0.1249 0.0999

3.89 3.32 3.89 4.44 3.78

    

106 105 105 105 105

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qe ¼ 15:04DH þ þ 6:253

(2)

This value is comparable with similar cases reported in the literature: 33.22 mg gmM1 for biosorption of lead(II) onto mansonia wood sawdust (Ofomaja, 2010), 0.289 mg gmM1 for cadmium(II) adsorption onto coconut copra meal (Ofomaja and Ho, 2007), and 10.60 mg gmM1 for lead(II) transfer onto tree fern (Ho, 2005). 3.3. Ion-exchange model The contribution of the ion-exchange mechanism to the absorption of mercury ions onto the dried biomass of V. parahaemolyticus PG02 can be assessed from the variation of initial mercury concentration (10e200 mg l1). Boyd et al. (1947) developed a rate equation, that considered rates of ion-exchange adsorption by organic zeolites as follows:

 S t 2:303

 logð1  FÞ ¼ 

(3)

where F is the fractional attainment of equilibrium, F¼qt/qe, and S (min1) is a constant. The application of this model was tested by Ofomaja and Ho (2007) and Ho (2005) for biosorption of lead(II). This model was used in this study for investigating the involvement of ion-exchange reactions in the biosorption of Hg2þ onto dried biomass of PG02. Fig. 3 shows a good fit of the experimental data by theoretical curves in the first 15 min of contact at all metal concentrations. After that, the experimental data do not follow the model predictions. Boyd et al. (1947) rate equation constants (S) at all initial mercury concentrations for this period are presented in Table 2, along with their standard error values. The decrease of the rate constant, S, at increasing metal concentration, is in accordance with the findings of Ofomaja (2010). Therefore, according to the Boyd et al. (1947) model, ionexchange mechanisms drove biosorption of mercury to dried biomass 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 kinetics of mercury biosorption was investigated by the use of several

15

initial mercury concentrations (10, 40, 100, and 200 mg l1 Hg2þ) at an 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 widely used relationships for modeling the sorption of a solute from a liquid solution. The linear form of Lagergren’s pseudo-first-order
lu and Arıca, 2008): rate equation is expressed as follows (Bayramog

logðqe  qt Þ ¼ log qe 

k1 t 2:303

(4)

where qt and qe are the amounts of Hg2þ adsorbed at time t (in minutes) and equilibrium (mol g1), respectively, and k1 is the rate constant of the pseudo-first-order adsorption process (min1). The values of k1, and qe are determined from the slope and intercept of the plot of log (qeqt) against t, respectively. The linear form of the pseudo-second-order kinetic model is given as Eq. (5) (Tunali Akar et al., 2012):

t 1 1 ¼ þ t qt k2 q2e qe

(5)

where k2 is the equilibrium rate constant for pseudo-second-order biosorption (g mol1 min1). Values of k2 and qe were calculated from 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 that the pseudo-second-order model describes the experimental data much better for the entire adsorption process (Fig. 4b). Khoramzadeh et al. (2013), however, reported that the first-order equation of Lagergren is applicable only over the initial 20e 30 min of the sorption process and not for the entire range of contact time investigated in this study. The kinetic parameters for both models along with their coefficients of correlations (R2) were tabulated in Table 2. The pseudosecond-order rate constant (k2) decrease from 8150.4 to 407.1 g mol1 min1 with an increase in the initial metal concentration from 10 to 200 mg l1 is consistent with the findings of other researchers (El-Sikaily et al., 2007; Ofomaja, 2010). The coefficient of correlation (R2) of the pseudo-second-order model (Eq. (5)) for all the studied concentrations (all above 0.992) was higher than that of the pseudo-first order model (Eq. (4)). In addition, pseudo-first-order theoretical qe values were not appropriate for description of the adsorption behavior. This suggests that Hg2þ biosorption by V. parahaemolyticus PG02 was not a first-order reaction. According to Sinha et al. (2012), if the experimental qe is not equal to the theoretical qe, the considered kinetic equation will not correspond to experimental data, even if the plot has a high coefficient of correlation. Analysis of the experimental data by the pseudo-second-order model suggests that the rate-limiting step was chemisorption (Lesmana et al., 2009). If the relationship between initial metal concentration and rate of adsorption is not linear (as seen in Table 2) then other factors such as intraparticle diffusion may limit the adsorption process (El-Sikaily et al., 2007).

3.5. Intraparticle diffusion Since the Eqs. (4) and (5) cannot replicate the diffusion mechanism satisfactorily, the intraparticle diffusion model was also evaluated (Gialamouidis et al., 2010):

qt ¼ kint t 0:5

(6)

2þ

Fig. 3. Ion-exchange biosorption kinetics of Hg onto dried Vibrio parahaemolyticus PG02 at various initial Hg2þ concentrations. Biosorbent concentration 1.0 g l1, pH 6.0, and temperature 35  C during 20 min of contact.

where kint is the intraparticle diffusion rate constant (mol g1 min0.5) and is the slope of the plot of qt versus t0.5

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Table 2 Pseudo-first order, Pseudo-second order, intraparticle diffusion, and ion-exchange model parameters for Hg2þ biosorption onto dried biomass of Vibrio parahaemolyticus PG2 at different initial concentrations. Ci (mg l1)

10 40 100 200 a b

qe,exp (mol g1)

4.53 1.76 4.18 7.34

   

105 104 104 104

Pseudo-first order model k1 (min1)

qe,model (mol g1)

0.061 0.049 0.053 0.047

2.4 8.6 2.2 4.2

   

105 105 104 104

Intraparticle diffusiona

Pseudo-second order model R2

k2 (g mol1 min1)

qe,model (mol g1)

0.917 0.834 0.839 0.903

8150.4 2165.8 755.2 407.1

4.6 1.8 4.2 7.3

   

105 104 104 104

R2

kint (mol g1 min0.5)

0.996 0.995 0.993 0.994

3.22 1.25 3.38 5.59

   

106 105 105 105

   

7.5  107 3.6  106 9  106 1.2  105

Ion-exchange modela,b R2

S (min1)

0.785 0.702 0.738 0.814

0.1252 0.1176 0.1157 0.1078

   

R2 0.002 0.000 0.002 0.003

0.998 0.999 0.998 0.996

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

(Fig. 5). The values of kint, at different metal concentrations are shown in Table 2 along with their standard error values. The results indicate that the intraparticle diffusion rate increased from 3.22  106 to 5.59  105 mol g1 min0.5 with increase of the initial mercury concentration from 10 to 200 mg l1. Ofomaja (2010) claimed that as initial metal concentration is increased, intraparticle diffusion becomes more important, because a higher initial metal concentration produces a stronger driving force for diffusion. According to Fig. 5, the overall line for each metal concentration does not pass through the origin. El-Sikaily et al. (2007) and Ofomaja (2010) reported that if the qt versus t0.5 plots do not pass through the origin, then the intraparticle diffusion is not the sole rate-limiting step. The poor R2 values for all initial mercury concentrations for intraparticle diffusion (Table 2) indicate that the process of mercury biosorption on dried biomass of V. parahaemolyticus PG02 occurred by surface biosorption and intraparticle diffusion.

This multi-linearity in the shape of the intraparticle diffusion plots has also been observed by Ofomaja (2010) in the biosorption of lead(II) onto mansonia wood sawdust and by Gialamouidis et al. (2010) for biosorption of Mn(II) by Pseudomonas sp., Staphylococcus xylosus, and Blakeslea trispora cells. 3.6. Biosorption isotherm In order to obtain information about the equilibrium, the maximum sorption capacity, and the affinity of the biosorbent to the specific metal ions, and also for comparing the capacity of different biosorbents with each other, Langmuir and Freundlich equilibrium isotherms have been traditionally used to evaluate sorption data. Equilibrium experiments were performed under predetermined conditions with initial mercury concentrations in the range of 10e350 mg l1 and 60 min of contact. The Langmuir model is based on a monolayer sorption on a homogeneous surface without interaction between adsorbed molecules (Plaza et al., 2011). Eq. (7) shows the linear form of the Langmuir isotherm (Sinha et al., 2012):

Ce 1 Ce þ ¼ qmax b qmax qe

(7)

where Ce is the equilibrium concentration (mol l1), qmax is the maximum metal uptake (mol g1) to form a complete monolayer on the surface and b is the Langmuir equilibrium constant (l mol1) that represents the affinity between the sorbent and adsorbate. The qmax and b values can be determined by plotting Ce/qe versus Ce. It should be emphasized that, in a strict theoretical sense, Ce in Eq. (7) must be expressed as the molar concentration (Liu, 2006).

Fig. 4. Plots of the (a) pseudo-first-order and (b) pseudo-second-order equations for the biosorption kinetics of Hg2þ onto dried Vibrio parahaemolyticus PG02 at various initial Hg2þ concentrations. Biosorbent concentration 1.0 g l1, pH 6.0, and temperature 35  C.

Fig. 5. Intraparticle diffusion mechanism for Hg2þ biosorption by dried Vibrio parahaemolyticus PG02 at various initial Hg2þ concentrations. Biosorbent concentration 1.0 g l1, pH 6.0, and temperature 35  C.

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17

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

Fig. 6. (a) Langmuir and (b) Freundlich isotherm plots for Hg2þ biosorption onto dried Vibrio parahaemolyticus PG02 cells at temperatures of 15 (A), 25 (-), 35 (:), and 45  C (). Biomass concentration 1.0 g l1, pH 6.0, and 60 min of contact.

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

1 log qe ¼ log kF þ log Ce n

(8)

where kF (mol g1) and n are the Freundlich constant and Freundlich exponent, respectively (Gialamouidis et al., 2010). The kF value is an indicator of the biosorption capacity and 1/n represents the surface heterogeneity and are calculated from the intercept and slope of a plot of log qe versus log Ce, respectively (Ertugay and Bayhan, 2010; Tunali Akar et al., 2012). Fig. 6 (a and b) illustrates the experimental equilibrium concentrations for mercury removal by dried biomass of V. parahaemolyticus PG02 at different temperatures, as predicted by the Langmuir and Freundlich models, respectively. It is obvious that the Langmuir isotherm model describes the experimental data better than the Freundlich model does.

Sorbent

Initial pH

qmax (mol g1)

Reference

Bacillus sp. Vibrio parahaemolyticus PG02 Chlamydomonas reinhardtii Ulva lactuca Bacillus subtilis Sugarcane Bagasse

4.5e6.0 6

3.94  105 9.63  104

Green-Ruiz (2006) This study

6.1  104

Tüzün et al. (2005)

5.5 5 4

4.22  10 3.56  104 1.78  104

Potamogeton natans

Natural

1.13  103

Spirulina platensis Walnut shell activated carbons Heat inactivated Phanerochaete chrysosporium

6 5

2.13  103 7.55  104

Zeroual et al. (2003) Wang et al. (2010) Khoramzadeh et al. (2013) Lacher and Smith (2002a) Cain et al. (2008) Zabihi et al. (2010)

5

8.58  104

Kacar et al. (2002)

6

4

The maximum sorption capacity, qmax, at pH 6 and a temperature of 35  C was 9.63  104 mol Hg2þ g1 biomass, which is high value in comparison with other reported sorbents for Hg2þ removal (Table 4). In the case of the Freundlich model, usually for a good adsorbent the value of n should be between 1 and 10. Larger values of n indicate a strong bond between biosorbent and heavy metal (Li et al., 2010). The values achieved for n were in the range of 1.53e1.78 for temperatures of 15e45  C. Although it seems that these values were not large, they were a little higher than those 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 different biosorbents. According to Table 3 the increase in temperature up to 35  C led to an increase in the sorption capacity from 5.08  104 to 9.63  104 mol g1, but higher temperatures reduced it to 8.86  104 mol g1. Presumably, increasing the temperature slightly increased the migration of Hg2þ ions into the inner structure of the biosorbent (Ofomaja, 2010). The significant decrease in biosorption capacity above 35  C was caused possibly by the destruction of surface binding ligands of PG02 (Gialamouidis et al., 2010). These results are in accordance with results reported by Green-Ruiz (2006).

3.7. Thermodynamic studies and the effect of temperature on Hg2þ biosorption In order to perform the thermodynamic studies, equilibrium experiments were repeated at four different temperaturesd15, 25, 35, and 45  Cdat a pH of 6. The free energy change of the adsorption, DG0 kJ mol1, as the fundamental criterion of spontaneity, was determined from Eq. (9) (Gialamouidis et al., 2010):

Table 3 Langmuir and Freundlich constants along with thermodynamic parameters at different temperatures for the biosorption of Hg2þ ions onto dried biomass of Vibrio parahaemolyticus PG2. Temperature

Langmuir constants 1

qmax (mol g 15 25 35 45 a



C C  C  C 

5.08 8.20 9.63 8.86

   

104 104 104 104

)

Thermodynamic parametersa

Freundlich constants b (l mol

1

3954.28 4875.78 10816.32 8244.73

)

2

R

kF (mol g

0.989 0.961 0.998 0.989

0.031 0.083 0.061 0.066

These parameters were provided  standard error values (Uncertainty).

1

)

n

R

DG0 (kJ mol1)

DH0 (kJ mol1)

DS0 (j mol1 K1)

1.63 1.53 1.78 1.68

0.914 0.879 0.950 0.902

29.45 30.99 34.07 34.46

18.97  1.44

167.93  4.8

2

18

DG0 ¼ RT ln K

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

References

(9)

where R is the gas constant (8.314 J mol1 K1), T is the absolute temperature, and K is the thermodynamic equilibrium constant without unit. The use of the Langmuir equilibrium constant, b, instead of the dimensionless thermodynamic equilibrium constant in Eq. (9) is incorrect, as reported by Milonji c (2007) and Liu (2009). The b constant with the units l mol1, can then K be made dimensionless by multiplying with 55.5, number of moles of water per liter of solution (Milonji c, 2007). The Langmuir equilibrium constant with units of l mol1 can be used directly for determination of DG0, without any modification, only if neutral or weak charge adsorbates or a dilute solution of charged adsorbate are used (Liu, 2009). The free energy change of the adsorption is:

DG0 ¼ DH 0  T DS0

(10)

where DH0 is the enthalpy change (kJ mol1) and DS0 is the entropy change (j mol1 K1), that were evaluated from the intercept and the slope of a plot of DG0 versus T, respectively (Table 3). The positive value of DH0 (18.97 kJ mol1) indicates that the sorption of Hg2þ ions by dried biomass of PG02 was endothermic (Lesmana et al., 2009; Liu, 2009). The increase in the values of qmax with temperature (Table 3) confirms the endothermic nature of the process. The positive value for enthalpy change also suggests a high probability that adsorption occurred mainly by chemisorption. However, heats of chemisorption are generally in the range of 80e200 kJ mol1, whereas the enthalpy change in this work was in the range of heats of condensation, from 2.1 to 20.9 kJ mol1. Therefore, the enthalpy change measured in this work suggests mercury adsorption onto dried biomass of PG02 occurred via physical adsorption (Liu, 2009; Gialamouidis et al., 2010). The free energy change values, DG0, were negative for all studied temperatures. This indicates that the biosorption process was spontaneous (Lesmana et al., 2009). The positive value of DS0 (167.93 j mol1 K1) indicates an increased randomness at the solid/solution interface during the adsorption of mercury onto biomass and some structural changes in the cell wall.

4. Conclusions The optimum conditions for mercury removal by V. parahaemolyticus PG02 were a pH 6 at 35  C. The ion-exchange mechanism was a predominant mechanism in the first 15 min of contact. Only the pseudo-second-order model described the experimental data for the entire range of contact time and the intraparticle diffusion was not the sole rate-limiting step. According to the Langmuir model, the qmax value was calculated to be 9.63  104 mol g1 at optimum conditions (pH value of 6 and temperature of 35  C). Thermodynamic parameters confirmed that the physical adsorption process was endothermic and spontaneous, with a high positive value for DS0. The results suggest the PG02 strain is a good adsorbent for mercury removal from aqueous solutions.

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


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