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Mechanism of enhanced Sb(V) removal from aqueous solution using chemically modified aerobic granules
Journal of Hazardous Materials 284 (2015) 43–49
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Journal of Hazardous Materials journal homepage: www.elsevier.com/locate/jhazmat
Mechanism of enhanced Sb(V) removal from aqueous solution using chemically modified aerobic granules Li Wang a,∗ , Chun-li Wan b , Yi Zhang b , Duu-Jong Lee c , Xiang Liu b , Xiao-feng Chen a , Joo-Hwa Tay d,∗ a
Center of Analysis and Measurement, Fudan University, Shanghai 200433, China Department of Environmental Science and Engineering, Fudan University, Shanghai 200433, China c Department of Chemical Engineering, National Taiwan University of Science and Technology, Taipei 106, Taiwan d Department of Civil Engineering, University of Calgary, Calgary, Canada b
h i g h l i g h t s • • • • •
Ionic strength significantly inhibits Sb(V) removal efficiency. Fe-modified granules almost have almost no affinity for cations. The maximum adsorption quantity is calculated to be 125 mg/g. Adsorption of antimony was spontaneous and endothermic. Outer and inner-sphere complexes could be formed stepwise during adsorption.
a r t i c l e
i n f o
Article history: Received 31 July 2014 Received in revised form 24 October 2014 Accepted 29 October 2014 Available online 4 November 2014 Keywords: Modification Adsorption Aerobic granules Mechanism
a b s t r a c t Sb(V) removal using Fe-modified aerobic granules was investigated. Increasing the biomass dosage improved the Sb(V) removal rate, but lowered the adsorption quantity; the optimal biomass concentration was 20 g/L (wet basis). Adsorption equilibrium was obtained at 2 h at 175 rpm; the adsorption quantity was 36.6 mg/g. NaCl and other salts inhibited Sb(V) adsorption on Fe-modified granules, and the mechanism possibly lied more with the anions. The adsorption isotherms were evaluated using the Langmuir, Freundlich, and Temkin models. The Langmuir model best described the adsorption process, and gave a maximum monolayer adsorption quantity of 125 mg/g. The H value for adsorption was 16.1 kJ/mol, indicating endothermicity, and the negative G values at various temperatures suggested spontaneous adsorption. Outer-sphere and inner-sphere complexations were involved in Sb(V) adsorption. © 2014 Elsevier B.V. All rights reserved.
1. Introduction Sb is a semi-metallic chemical element, and its compounds are widely used in flame retardants, alloys, batteries, and therapeutic agents against protozoan diseases [1,2]. Uncontrolled discharge of Sb-containing wastewater by mining and smelting industries causes severe pollution in the receiving environment. Well water has been polluted by Sb in wastewater at concentrations of 24.02–42.03 mg/L [3]. Sb is a toxic heavy metal with potential carcinogenic effects [4], therefore, Sb and its compounds are serious
∗ Corresponding authors. Tel.: +86 13 564198402/21 65643065; fax: +86 21 65643065. E-mail addresses: [email protected] (L. Wang), [email protected] (J.-H. Tay). http://dx.doi.org/10.1016/j.jhazmat.2014.10.041 0304-3894/© 2014 Elsevier B.V. All rights reserved.
threats to human health, and have been listed as priority pollutants by the European Union and the US Environmental Protection Agency; the maximum permissible concentrations of Sb in drinking water were set at 6 and 10 g/L, respectively, by these two agencies [5]. The limit imposed by the World Health Organization is 5 g/L [6]. Sb(III) and Sb(V) are the main oxidation states of Sb in the environment, and Sb(III) is 10 times more toxic than Sb(V) [2]. In the pH range 2–11, Sb(V) exists as a negatively charged oxyanion, Sb(OH)6 − , and Sb(III) exists as an uncharged neutral complex, Sb(OH)3 [7,8]. Much attention has been paid to the removal of Sb(III) [9–11], but there have been fewer Sb(V) studies. Sb has been observed to change its valence in response to changes in the redox potential of the surrounding environment [12,13], therefore, removal of Sb(V) is of equal importance. Various methods have
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been used to remove Sb from aqueous media such as chemical precipitation, ion exchange, reverse osmosis, solvent extraction, and adsorption [9]. Aerobic granules are novel microbial aggregates, cultivated in sequencing batch reactors (SBRs) from activated sludge inoculation [14]. Aerobic granules have been investigated as biosorbents for the removal of heavy metals, because the granules have high physical strength, high specific gravity, and good settleability [15,16]. However, most studies so far have been conducted using metal cations, based on removal resulting from electrostatic attraction between negatively charged granules and positively charged cations. This mechanism is less efficient if the ions to be removed bear negative charges, such as Sb(OH)6 − . A possible solution to this problem is to perform surface modification of the granules to change the nature of the surface charge and improve the removal of oxyanionic heavy metals. To date, only a few studies on the adsorption of oxyanions on modified aerobic granules have been carried out [17,18]. Because of cost and secondary pollution considerations, modification using common non-toxic inorganic salts is of particular interest; for example, a preliminary study of adsorption of Sb(V) on Fe-modified granules has been reported [19], in which the effects of pH, adsorption kinetics, and characteristics of the granules were explored. However, other important factors, including biomass dosage, shaker speed, ionic strength, and interfering ions, which could affect the performances of aerobic granules in Sb(V) adsorption, have not yet been fully investigated. Modification with FeCl3 changes the surface properties of the granules, and makes adsorption of Sb(OH)6 − on Fe-modified aerobic granules complicated, which has not been investigated in sufficient detail. These aspects are specifically addressed in this study. The objectives were (I) identification of factors affecting the adsorption capacity of granules; (II) further study of the adsorption isotherms and thermodynamics, to determine the adsorption characteristics of the granules; and (III) investigation of the adsorption mechanism of the negatively charged Sb(OH)6 − on Fe-modified aerobic granules. In this paper, we describe the macroscopic adsorption characteristics and microscopic removal mechanism. 2. Material and methods 2.1. Preparation and modification of aerobic granules Aerobic granules were cultivated in a column-type SBR, using a previously reported procedure [20]. The modification process was performed by mixing fresh granules with FeCl3 solution, as described in detail elsewhere [19]. Fe-modified granules were stored in tap water at 4 ◦ C for future use.
effect of ionic strength on the adsorption capacity. In addition, the interference effects of Cu2+ , Ni2+ , and Zn2+ at concentrations of 20, 40, and 60 mg/L on Sb(V) adsorption by granules were investigated. The test time was 300 min, and samples were collected at fixed time intervals for analysis. Each sample was centrifuged at 9000 × g for 5 min, and the Sb concentration in the supernatant was determined. The specific adsorption quantity, qt (mg Sb/g dry granules) and removal rate, r (% removal), are expressed as follows: qt = (C0 − Ce )v/m and r = (C0 − Ce )/C0 , where C0 and Ce are the initial and equilibrium concentrations, respectively, of Sb (mg/L), v is the volume of the solution (L), and m is the dry weight of modified granules (g). All experiments in this study were performed in triplicate. Metal-free and granules-free sample have been included as blank control. The results showed that the Sb(V) adsorbed onto the flasks was negligible, so was the Sb-release by granules. 2.4. Analytical methods The Sb(V) concentration was determined using inductively coupled plasma atomic emission spectroscopy (P-4010, Japan). The concentrations of Cu2+ , Ni2+ , and Zn2+ were determined using flame atomic absorption spectroscopy (Z-5000, Japan). The mass fraction of Fe in the granules was determined by dissolution and ICP-AES. 3. Results and discussion 3.1. Adsorption studies 3.1.1. Effect of biomass dosage The effect of biomass dosage on the adsorption of Sb(V) by Femodified aerobic granules is shown in Fig. 1. The Sb(V) removal rate increased with increasing biomass concentration from 2 to 30 g/L (wet basis); it reached a maximum of 99.3% at 20 g/L, and then remained almost unchanged. An increase in the biomass concentration results in a large surface area and more adsorption sites, which could bind more Sb(V) and improve the removal rate. However, the adsorption quantity decreased from 52.4 to 15.1 mg/g when the biomass concentration increased from 2 to 30 g/L. Liu et al. found that the ratio of initial Zn(II) concentration to initial granule concentration was closely associated with the biosorption quantity, and the ratio would better reflect the real driving force for metal biosorption by micro-organisms [21]. In present study, increasing biomass concentration at fixed initial Sb(V) concentration reduced the ratio, and accordingly weakened the driving force
2.2. Chemicals A standard stock solution of Sb(V) (1000 mg/L) was prepared by dissolving KSb(OH)6 in deionized water, and stock solutions of Cu2+ , Ni2+ , and Zn2+ (1000 mg/L) were prepared by dissolving CuSO4 , NiSO4 , and Zn(CH3 OOH)2 , respectively, in deionized water. The stock solutions were diluted to the required concentrations for the experiments, and the solution pHs were adjusted by adding 0.1 M HCl or NaOH. 2.3. Adsorption experiments The effects on adsorption of four major factors were studied. Different concentrations of modified granules, from 2 to 30 g/L (wet basis), were placed in flasks on a rotary shaker, and the shaking speed was varied from 0 to 175 rpm to investigate the effect of mass transfer in the bulk liquid. NaCl concentrations of 0.1 M and 0.3 M were supplied as background electrolytes, to evaluate the
Fig. 1. Effect of biomass dosage on the adsorption of Sb(V) on Fe-modified granules (T: 35 ◦ C, shaking speed: 175 rpm, C0 : 20 mg/L, pH: 3.4, contact time: 300 min).
L. Wang et al. / Journal of Hazardous Materials 284 (2015) 43–49
Fig. 2. Effect of shaking speed on the adsorption of Sb(V) on Fe-modified granules (T: 35 ◦ C, biomass dosage: 0.4 g/L, C0 : 20 mg/L, pH: 3.4).
for Sb(V) diffusion into the inside of the granules, which could lead to increasing unsaturation of adsorption sites per unit of biomass. When all Sb(V) was adsorbed on the granules, the contribution of additional biomass would be insignificant [22], and the specific adsorption quantity was therefore negatively correlated with the biomass dosage. These results suggest that the use of excessive amounts of biomass is uneconomic; a biomass dosage of 20 g/L (wet basis) gave the maximum efficiency in this study. 3.1.2. Effect of shaking speed External mixing of bulk liquid was expressed by shaking speed. Increasing the shaking speed increases the fluid shear stress, which may influence the boundary conditions and adsorption efficiency on the granule surface. Various shaking speeds were used to investigate the influence of shear stress, and the changes in adsorption quantity with time are shown in Fig. 2. The difference between adsorption efficiency at shaking speeds of 0 and 50 rpm was insignificant, because electrostatic interactions and film diffusion control the process under low shear, and a longer time was needed to reach adsorption equilibrium. When the shaking speed was increased to 100 rpm, the adsorption capacity of the Fe-modified granules improved significantly, and the adsorption quantity rose by 20% at 5 h. The increasing turbulence in the bulk liquid significantly improved mixing of the granules with soluble Sb(V), and the stronger driving force facilitated diffusion of Sb into the interiors of the granules. However, the use of a high speed, i.e., 175 rpm, did not improve the adsorption efficiency significantly, but a shorter time, i.e., 2 h, was needed to reach the adsorption equilibrium, compared with other speeds. In this study, a shaking speed of 175 rpm was therefore, used. 3.1.3. Effect of ionic strength NaCl was used to simulate various ionic strengths in the aqueous media. The adsorption quantity of Sb(V) on Fe-modified granules with time at different NaCl concentrations are shown in Fig. 3. A higher ionic strength significantly decreased the adsorption capacity, by 25% and 45% at 5 h with NaCl concentrations of 0.1 M and 0.3 M, respectively. There are several causes of this phenomenon. One reason is that at high concentrations, Cl− anion compete with Sb(OH)6 − for the available sites on the granule surface, and prevent Sb(V) approaching the granules. Second, Sb(V) exists mainly as highly polymerised hydroxy–chloro complexes or as colloidal hydrous oxides in the presence of high Cl− concentrations [13], so the original Sb(V) species may change. In addition, KSb(OH)6 is a
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Fig. 3. Effect of ionic strength on the adsorption of Sb(V) on Fe-modified granules (T: 35 ◦ C, biomass dosage: 0.4 g/L, C0 : 20 mg/L, pH: 3.4, speed: 175 rpm).
common reagent for detecting Na+ in aqueous media, as follows [23]: NaCl + KSb(OH)6 → NaSb(OH)6 ↓ + KCl
(1)
The formation of a precipitate from these two reagents also changes the ionic state of Sb(V). The reasons stated above could individually, or in combination, lower the adsorption capacity of Fe-modified aerobic granules for Sb(V) on addition of NaCl as a background electrolyte. 3.1.4. Effect of interfering ions In the mining and exploitation of Sb, a large number of accompanying toxic cations, e.g., Cu2+ , Ni2+ , and Zn2+ , are commonly included in the Sb-containing wastewater stream produced [24]. Various concentrations of these cations were mixed with Sb(V) (20 mg/L) at pH 3.4 for 5 h. The adsorption quantity as a function of cation concentrations are shown in Fig. 4. The addition of interfering ions clearly inhibited adsorption of Sb(V) on the Fe-modified granules. The concentrations of Cu2+ , Zn2+ , and Ni2+ before and after adsorption were determined to investigate the cation-binding capacity of the Fe-modified granules. The results indicate that the concentrations of all the cations were almost unchanged after mixing with the granules.
Fig. 4. Effect of interfering ions on the adsorption of Sb(V) on Fe-modified granules (T: 35 ◦ C, biomass dosage: 0.4 g/L, C0 : 20 mg/L, pH: 3.4, speed: 175 rpm).
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Modification with FeCl3 made the surface of the aerobic granules positively charged, and prevented the cations from approaching the adsorbent. Cu2+ , Zn2+ , and Ni2+ , therefore, did not compete with Sb(OH)6 − for adsorption sites. Consequently, The inhibitory effect might be attributed to the corresponding anions by competing for adsorption sites with Sb(OH)6 − . Xi et al. found that the Sb(V) adsorption on bentonite was significantly affected in the presence of three competitive anions, and the effect was more evident in the case of PO4 3− and SO4 2− as compared to NO3 − [25]. The possible accumulation or precipitation of competitive anions on the surface of adsorbent could increase the negative surface charge and electrostatic repulsion with Sb(V), which might be the main reason of the inhibitory effect. In this study, CuSO4 , NiSO4 , and Zn(CH3 OOH)2 were used to generate interfering cations, therefore, the corresponding anions were SO4 2− , SO4 2− , and CH3 COO− , at molar concentrations of 0.34, 0.31, and 0.61 mmol/L, respectively, when the mass concentration of cations was all fixed at 20 mg/L. SO4 2− is an inorganic ion, while CH3 COO− is an organic weak acid anion, which could form CH3 COOH at acidic condition. The possible surface complexation between organic weak acid and Sb(V) could change the original species of Sb(OH)6 − , and reduce the adsorption capacity of the granules for Sb(V) [26]. Generally, the existence of anions could contribute to the decrease of Sb(V) adsorption onto the Fe-modified granules.
Table 1 Isotherm parameters for the adsorption of Sb(V) on Fe-modified granules. Langmuir
qm (mg/g) 125
b (L/mg) 0.056
R2 0.987
Freundlich
kF (mg/g) 27.1
n 3.9
R2 0.963
Temkin
B 12.2
A (L/mg) 18.2
R2 0.920
3.2. Isotherm analysis The adsorption of Sb(V) on Fe-modified granules was performed until equilibrium was reached, and the data were analyzed using three isotherm models, i.e., the Langmuir, Freundlich, and Temkin models. The Langmuir isotherm model assumes that the adsorbent surface has homogeneous sites, and monolayer adsorption occurs without interactions between adsorbed molecules [27]. The linear form of the Langmuir isotherm model can be expressed as [28]: Ce 1 Ce = + qm qe (bqm )
(2)
where Ce is the equilibrium concentration of Sb(V) (mg/L), qe is the equilibrium adsorption quantity (mg/g), qm is the maximum adsorption quantity (mg/g), and b is the Langmuir equilibrium constant (L/mg). The Freundlich isotherm model is used to describe adsorption on heterogeneous surface through multilayer adsorption [20], which can be described as [29]: lnqe =
1 n ln C e + lnkF
(3)
where kF is the Freundlich constant, representing the adsorption capacity (mg/g), and 1/n is related to the adsorption intensity of the adsorbent. The Temkin isotherm model assumes that the heat of adsorption of the molecules decreases linearly with coverage, because of the interactions between the adsorbent and adsorbate, and a uniform distribution of binding energies characterizes the adsorption up to a maximum energy; this model can be expressed as [30]: qe = BlnC e +BlnA
(4)
where B is related to the heat of adsorption and A represents the equilibrium binding constant (L/mg), corresponding to the maximum binding energy. Adsorption isotherms are indispensable in investigating the nature of the interactions between adsorbents and adsorbates [22]. The equilibrium data were fitted to the Langmuir, Freundlich, and Temkin isotherm models; the results are shown in Table 1 and Fig. 5. The order of the correlation coefficients, R2 , for the biosorption
Fig. 5. Adsorption isotherm of Sb(V) on Fe-modified granules by different models.
of Sb(V) is Langmuir > Freundlich > Temkin, therefore, the experimental equilibrium adsorption data are best fitted by the Langmuir model. The maximum adsorption quantity and Langmuir constant were calculated to be 125 mg/g and 0.056 L/mg, respectively. This indicated that adsorption of Sb(V) on the Fe-modified granules mainly occurred in a monolayer, without interactions between adsorbed molecules; these results are similar to those reported for biosorption of Cu2+ on modified Penicillium biomass [31]. However, based on the previous studies, the removal rate of Sb(V) in aqueous solution by original and Fe-modified granules were 14% and 99% at pH 3.4, respectively [19]. This suggested that the calculated maximum adsorption quantity of 125 mg/g in the present study could be the result of combined action by original granules and Fe-modification. 3.3. Thermodynamics analysis In thermodynamics, it is assumed that energy cannot be gained or lost in an isolated system, and the entropy change represents the driving force [32]. The changes in entropy (S) and enthalpy (H) are important thermodynamics parameters for the identification of spontaneous process [33]. The thermodynamic parameters S, H, and the Gibbs free energy (G) can be calculated using the following equations [27]: G = −RT lnK K=
qe Ce
G = H − TS
(5) (6) (7)
where R is the ideal gas constant [8.314 × 10−3 kJ/(mol K)], T is the absolute temperature (K), and K is the distribution coefficient, which is the ratio of the equilibrium adsorption quantity to the equilibrium concentration of Sb(V).
L. Wang et al. / Journal of Hazardous Materials 284 (2015) 43–49
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Table 2 Thermodynamic parameters for the adsorption of Sb(V) on Fe-modified granules. T (K)
G kJ/(mol K)
H kJ/(mol K)
S kJ/(mol K)
283 298 308 318
−4.87 −5.98 −6.72 −7.46
16.1
74.1
chemisorption [33]. The positive value of S suggests an increase in randomness at the solid/liquid interface during adsorption of Sb(V) on the Fe-modified granules. The values of G, H, and S are shown in Table 2. 3.4. Test of acid-soluble Fe in Fe-modified granules
Fig. 6. Plots of lnK versus 1/T for adsorption of Sb(V) on Fe-modified granules.
The final equation can be written as: lnK =
S H − R (RT )
(8)
The entropy change (S) and enthalpy change (H) can be calculated from the slope and intercept of a plot of 1/T versus lnK, as shown in Fig. 6. The Gibbs free energy (G) can be determined from H, S, and T using Eq. (7). The negative values of G indicate that adsorption of Sb(V) on Fe-modified granules is spontaneous in the temperature range of 283–318 K. The absolute value of G increases with increasing temperature, which suggests that adsorption occurs easily at high temperatures. The positive value of H indicates that Sb(V) adsorption on Fe-modified granules is endothermic, which agrees with the calculated results for the Gibbs free energy. Ions that dissolve well in water need to be dehydrated to be adsorbed on an adsorbent. When the dehydration process requires energy exceeding the exothermicity of ion attachment to the surface, a high temperature provides more energy for dehydration of Sb(V) and more activation energy for monolayer
Various initial concentrations of Sb(V) were used in the adsorption experiments. After adsorption, the Sb-loaded and the original granules were firstly dissolved in nitrohydrochloric acid, and precipitations were found in all samples except the granules without Sb dosage. The pretreatment was followed by drying, carbonization, and ashing in muffle furnace. The residues were further immersed in nitrohydrochloric acid, and the sample unloaded with Sb could be dissolved completely, leaving clear acid solution. However, insoluble precipitation could be observed in all the other samples. The concentrations of Fe in the supernatant were further determined using ICP-AES. The mass fractions of acid-soluble Fe in the Sbbearing granules were calculated and shown in Fig. 7. Before adsorption, the concentration of Fe in the modified granules was 5.9% (w/w). With increasing initial Sb(V) concentration, the concentration of acid-soluble Fe in the granules decreased. For initial concentrations of 300 and 500 mg/L, the amounts of acidsoluble Fe decreased to 2.4% and 2.3%. This could be attributed to the possible formation of stable complexes, which cannot be easily broken up by acid. In addition, the higher the initial concentration of Sb(V) is, the more Fe on the granules could be participating in adsorption, hence the lower Fe concentration in the supernatant. However, some Fe in the inner layers of the granules might not be
Fig. 7. Concentration of acid-soluble iron in Fe-modified granules.
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accessible to Sb(V), which might be why there was still a certain level of acid-soluble Fe under high Sb(V) concentration. 3.5. Adsorption mechanism
protonation, and Fe OH is positively charged, whereas deprotonation occurs at high pH, and Fe OH is negatively charged. The reaction equations can be expressed as: Fe OH + H+ →
It has been reported that the macroscopic evidence of formation of inner-sphere complexes is the adsorption process’ insensitivity to changes of ionic strength; otherwise, it was more possible that out-sphere complexes was formed [34]. However, Xi et al. studied the adsorption of Sb(V) on kaolinite, and found that different ionic strengths had different effects on adsorption, which suggested that both inner-sphere and outer-sphere complexes were formed [34]. Similarly the influence of ionic strength in this study was found to be complicated, therefore the adsorption mechanism of Sb(V) on the Fe-modified granules could also be multiple and complicated. In a previous study, a similar modification of biomass was performed using FeCl3 , and it was found that Fe(III) existed in the form of Fe OH on biomass surface [35,36]. Therefore, it could be inferred that in present work, hydrolysis of Fe(III) in the solution could lead to formation of Fe(OH)3 , which were then stably attached onto the granules. Fe-modified granules thus have a large amount of Fe-hydroxyl groups, i.e., Fe OH, on the surface. The speciation of Fe OH depends on solution pH. A low pH leads to
−
Fe OH + OH →
Fe OH2 + −
Fe O + H2 O
(9) (10)
On the other hand, some researchers suggested that adsorption of Sb(V) onto hydrous ferric oxide (HFO) can be described by an inner-sphere surface complex model, the reaction of which is as follow [37]: Fe OH + Sb(OH)6 − = FeOSb(OH)5 − + H2 O
(11)
Similarly, the formation of complex between OH and Sb(OH)6 – by dehydration could contribute significantly to Sb(V) removal in present study. In theory, one Fe atom could have more than one OH group attached to it, and could bind more than one Sb atom, which was the reason why Fe-modified granules had a high adsorption capacity for Sb(V). The mechanism of adsorption of Sb(V) on Fe-modified granules therefore could be inferred, based on the results of our experiments and previous reports [35,36,38]. Under acidic condition, Fe-modified granules tend to be protonated to Fe OH2 + ·Sb(OH)6 − can therefore be
Fig. 8. The four steps of adsorption of Sb(V) on Fe-modified granules.
L. Wang et al. / Journal of Hazardous Materials 284 (2015) 43–49
easily attracted by positively charged granules through electrostatic interaction and approach the granules surface. Outer-sphere complexes [ Fe OH2 + ·Sb(OH)6 − ] can be formed through ion exchange, followed by formation of more stable chemical bonds ( Fe O Sb), and inner-sphere complexes are finally formed. The four possible steps in adsorption of Sb(V) on Fe-modified aerobic granules are shown in Fig. 8. 4. Conclusion Fe-modified aerobic granules were excellent biosorbents for the removal of Sb(V). Adsorption batch experiments suggested increasing the biomass dosage increased the number of adsorption sites and further enhanced the removal rate of Sb(V), but decreased the specific adsorption capacity. Increasing the shaking speed improved the adsorption capacity of the biosorbent and shortened the equilibrium adsorption time. The adsorption capacity of the Fe-modified granules was found to be inhibited significantly by addition of foreign ions, and the Fe-modified granules had almost no affinity for cations in aqueous media. The Langmuir model best fitted the equilibrium data, with a monolayer adsorption quantity of 125 mg/g. The adsorption of Sb(V) on the Femodified granules was found to be spontaneous and endothermic. Surface protonation, electrostatic interactions, ion exchange, and surface complexation were inferred to be the main mechanisms of adsorption of Sb(V) on Fe-modified granules. References [1] F. Sun, F. Wu, H. Liao, B. Xing, Biosorption of antimony(V) by freshwater cyanobacteria Microcystis biomass: chemical modification and biosorption mechanisms, Chem. Eng. J. 171 (2011) 1082–1090. [2] F. Wu, F. Sun, S. Wu, Y. Yan, B. Xing, Removal of antimony(III) from aqueous solution by freshwater cyanobacteria Microcystis biomass, Chem. Eng. J. 183 (2012) 172–179. [3] M. He, J. Yang, Effects of different forms of antimony on rice during the period of germination and growth and antimony concentration in rice tissue, Sci. Total Environ. 243 (1999) 149–155. [4] Z. Wu, M. He, X. Guo, R. Zhou, Removal of antimony(III) and antimony(V) from drinking water by ferric chloride coagulation: competing ion effect and the mechanism analysis, Sep. Purif. Technol. 76 (2010) 184–190. [5] W. Xu, H. Wang, R. Liu, X. Zhao, J. Qu, The mechanism of antimony(III) removal and its reactions on the surfaces of Fe–Mn binary oxide, J. Colloid Interface Sci. 363 (2011) 320–326. [6] X.J. Guo, Z.J. Wu, M.C. He, Removal of antimony(V) and antimony(III) from drinking water by coagulation–flocculation–sedimentation (C–F–S), Water Res. 43 (2009) 4327–4335. [7] A.K. Leuz, C.A. Johnson, Oxidation of Sb(III) to Sb(V) by O2 and H2 O2 in aqueous solutions, Geochim. Cosmochim. Acta 69 (2005) 1165–1172. [8] F. Quentel, M. Filella, C. Elleouet, C.L. Madec, Kinetic studies on Sb(III) oxidation by hydrogen peroxide in aqueous solution, Environ. Sci. Technol. 38 (2004) 2843–2848. [9] M.A. Salam, R.M. Mohamed, Removal of antimony(III) by multi-walled carbon nanotubes from model solution and environmental samples, Chem. Eng. Res. Des. 91 (2013) 1352–1360. [10] M. Iqbal, A. Saeed, R.G. Edyvean, Bioremoval of antimony(III) from contaminated water using several plant wastes: optimization of batch and dynamic flow conditions for sorption by green bean husk (Vigna radiata), Chem. Eng. J. 225 (2013) 192–201. [11] Y. Leng, W. Guo, S. Su, C. Yi, L. Xing, Removal of antimony(III) from aqueous solution by graphene as an adsorbent, Chem. Eng. J. 211 (2012) 406–411. [12] N. Belzile, Y.W. Chen, Z.J. Wang, Oxidation of antimony(III) by amorphous iron and manganese oxyhydroxides, Chem. Geo. 174 (2001) 379–387.
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[13] M. Filella, N. Belzile, Y.W. Chen, Antimony in the environment: a review focused on natural waters: II. Relevant solution chemistry, Earth-Sci. Rev. 59 (2002) 265–285. [14] D.W. Gao, L. Liu, H. Liang, W.M. Wu, Comparison of four enhancement strategies for aerobic granulation in sequencing batch reactors, J. Hazard. Mater. 186 (2011) 320–327. [15] Y. Liu, H. Xu, Equilibrium, thermodynamics, and mechanisms of Ni2+ biosorption by aerobic granules, Biochem. Eng. J. 35 (2007) 174–182. [16] L. Yao, Z.F. Ye, M.P. Tong, P. Lai, J.R. Ni, Removal of Cr3+ from aqueous solution by biosorption with aerobic granules, J. Hazard. Mater. 165 (2009) 250–255. [17] X.F. Sun, C. Liu, Y. Ma, S.G. Wang, B.Y. Gao, X.M. Li, Enhanced Cu(II) and Cr(VI) biosorption capacity on poly (ethylenimine) grafted aerobic granular sludge, Colloid Surf. B 82 (2011) 456–462. [18] X.F. Sun, Y. Ma, X.W. Liu, S.G. Wang, B.Y. Gao, X.M. Li, Sorption and detoxification of chromium(VI) by aerobic granules functionalized with polyethylenimine, Water Res. 44 (2010) 2517–2524. [19] L. Wang, C. Wan, D.J. Lee, X. Liu, Y. Zhang, X. Chen, J.H. Tay, Biosorption of antimony(V) onto Fe(III)-treated aerobic granules, Bioresour. Technol. 158 (2014) 351–354. [20] L. Wang, C. Wan, D.J. Lee, J.H. Tay, X. Chen, X. Liu, Y. Zhang, Adsorption–desorption of strontium from waters using aerobic granules, J. Taiwan Inst. Chem. Eng. 44 (2013) 454–457. [21] Y. Liu, S.F. Yang, S.F. Tan, Y.M. Lin, J.H. Tay, Aerobic granules: a novel zinc biosorbent, Lett. Appl. Microbiol. 35 (2002) 548–551. [22] F. Sun, W.L. Sun, H.M. Sun, J.R. Ni, Biosorption behavior and mechanism of beryllium from aqueous solution by aerobic granule, Chem. Eng. J. 172 (2011) 783–791. [23] X.Z. Cao, M.L. Zhao, D.W. Wang, The influences of EMP irradiation of rat hypothalamus in vitro on intracellular LDH and LDH, AST, CHE, K+ , and Na+ in culture medium, Acta Biophy. Sin. 18 (2002) 302–306. [24] X.Y. Shan, S.Q. Xu, The analysis and assessment on the pollution condition of heavy metals in the soil around the Qinglong Dachang antimony mining area in Guizhou province, J. Guizhou Univ. (Nat. Sci.) 28 (2011) 132–135. [25] J.H. Xi, M.C. He, C.Y. Lin, Adsorption of antimony(III) and antimony(V) on bentonite: kinetics, thermodynamics and anion competition, Microchem. J. 97 (2011) 85–91. [26] V.K. Mittal, S. Bera, S. Narasimhan, S. Velmurugan, Adsorption behavior of antimony(III) oxyanions on magnetite surface in aqueous organic acid environment, Appl. Surf. Sci. 266 (2013) 272–279. [27] D. Ozdes, C. Duran, H.B. Senturk, Adsorptive removal of Cd(II) and Pb(II) ions from aqueous solutions by using Turkish illitic clay, J. Environ. Manage. 92 (2011) 3082–3090. [28] I. Langmuir, The adsorption of gases on plane surfaces of glass, mica, and platinum, J. Am. Chem. Soc. 40 (1918) 1361–1403. [29] H. Freundlich, Über Die Adsorption in Lösungen, Engelmann, Leipzig, 1906. [30] M. Temkin, V. Pyzhev, Kinetics of ammonia synthesis on promoted iron catalysts, Acta Physiochim. U.R.S.S. 12 (1940) 217–222. [31] X. Zhang, H.J. Su, T.W. Tan, G. Xiao, Study of thermodynamics and dynamics of removing Cu(II) by biosorption membrane of Penicillium biomass, J. Hazard. Mater. 193 (2011) 1–9. [32] K.V. Kumar, A. Kumaran, Removal of methylene blue by mango seed kernel powder, Biochem. Eng. J. 27 (2005) 83–93. [33] M.V. Subbaiah, G. Yuvaraja, Y. Vijaya, A. Krishnaiah, Equilibrium, kinetic, and thermodynamic studies on biosorption of Pb(II) and Cd(II) from aqueous solution by fungus (Trametes versicolor) biomass, J. Taiwan Inst. Chem. Eng. 42 (2011) 965–971. [34] J. Xi, M. He, C. Lin, Adsorption of antimony(V) on kaolinite as a function of pH, ionic strength and humic acid, Environ. Earth Sci. 60 (2010) 715–722. [35] M. Aryal, M. Ziagova, M. Liakopoulou-Kyriakides, Study on arsenic biosorption using Fe(III)-treated biomass of Staphylococcus xylosus, Chem. Eng. J. 162 (2010) 178–185. [36] M. Aryal, M. Ziagova, M. Liakopoulou-Kyriakides, Comparison of Cr(VI) and As(V) removal in single and binary mixtures with Fe(III)-treated Staphylococcus xylosus biomass: thermodynamic studies, Chem. Eng. J. 169 (2011) 100–106. [37] X.J. Guo, Z.J. Wu, M.C. He, Removal of antimony(V) and antimony(III) from drinking water by coagulation–flocculation–sedimentation (C–F–S), Water Res. 43 (2009) 4327–4335. [38] A.K. Leuz, H. Mönch, C.A. Johnson, Sorption of Sb(III) and Sb(V) to goethite: influence on Sb(III) oxidation and mobilization, Environ. Sci. Technol. 40 (2006) 7277–7282.
Contents lists available at ScienceDirect
Journal of Hazardous Materials journal homepage: www.elsevier.com/locate/jhazmat
Mechanism of enhanced Sb(V) removal from aqueous solution using chemically modified aerobic granules Li Wang a,∗ , Chun-li Wan b , Yi Zhang b , Duu-Jong Lee c , Xiang Liu b , Xiao-feng Chen a , Joo-Hwa Tay d,∗ a
Center of Analysis and Measurement, Fudan University, Shanghai 200433, China Department of Environmental Science and Engineering, Fudan University, Shanghai 200433, China c Department of Chemical Engineering, National Taiwan University of Science and Technology, Taipei 106, Taiwan d Department of Civil Engineering, University of Calgary, Calgary, Canada b
h i g h l i g h t s • • • • •
Ionic strength significantly inhibits Sb(V) removal efficiency. Fe-modified granules almost have almost no affinity for cations. The maximum adsorption quantity is calculated to be 125 mg/g. Adsorption of antimony was spontaneous and endothermic. Outer and inner-sphere complexes could be formed stepwise during adsorption.
a r t i c l e
i n f o
Article history: Received 31 July 2014 Received in revised form 24 October 2014 Accepted 29 October 2014 Available online 4 November 2014 Keywords: Modification Adsorption Aerobic granules Mechanism
a b s t r a c t Sb(V) removal using Fe-modified aerobic granules was investigated. Increasing the biomass dosage improved the Sb(V) removal rate, but lowered the adsorption quantity; the optimal biomass concentration was 20 g/L (wet basis). Adsorption equilibrium was obtained at 2 h at 175 rpm; the adsorption quantity was 36.6 mg/g. NaCl and other salts inhibited Sb(V) adsorption on Fe-modified granules, and the mechanism possibly lied more with the anions. The adsorption isotherms were evaluated using the Langmuir, Freundlich, and Temkin models. The Langmuir model best described the adsorption process, and gave a maximum monolayer adsorption quantity of 125 mg/g. The H value for adsorption was 16.1 kJ/mol, indicating endothermicity, and the negative G values at various temperatures suggested spontaneous adsorption. Outer-sphere and inner-sphere complexations were involved in Sb(V) adsorption. © 2014 Elsevier B.V. All rights reserved.
1. Introduction Sb is a semi-metallic chemical element, and its compounds are widely used in flame retardants, alloys, batteries, and therapeutic agents against protozoan diseases [1,2]. Uncontrolled discharge of Sb-containing wastewater by mining and smelting industries causes severe pollution in the receiving environment. Well water has been polluted by Sb in wastewater at concentrations of 24.02–42.03 mg/L [3]. Sb is a toxic heavy metal with potential carcinogenic effects [4], therefore, Sb and its compounds are serious
∗ Corresponding authors. Tel.: +86 13 564198402/21 65643065; fax: +86 21 65643065. E-mail addresses: [email protected] (L. Wang), [email protected] (J.-H. Tay). http://dx.doi.org/10.1016/j.jhazmat.2014.10.041 0304-3894/© 2014 Elsevier B.V. All rights reserved.
threats to human health, and have been listed as priority pollutants by the European Union and the US Environmental Protection Agency; the maximum permissible concentrations of Sb in drinking water were set at 6 and 10 g/L, respectively, by these two agencies [5]. The limit imposed by the World Health Organization is 5 g/L [6]. Sb(III) and Sb(V) are the main oxidation states of Sb in the environment, and Sb(III) is 10 times more toxic than Sb(V) [2]. In the pH range 2–11, Sb(V) exists as a negatively charged oxyanion, Sb(OH)6 − , and Sb(III) exists as an uncharged neutral complex, Sb(OH)3 [7,8]. Much attention has been paid to the removal of Sb(III) [9–11], but there have been fewer Sb(V) studies. Sb has been observed to change its valence in response to changes in the redox potential of the surrounding environment [12,13], therefore, removal of Sb(V) is of equal importance. Various methods have
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been used to remove Sb from aqueous media such as chemical precipitation, ion exchange, reverse osmosis, solvent extraction, and adsorption [9]. Aerobic granules are novel microbial aggregates, cultivated in sequencing batch reactors (SBRs) from activated sludge inoculation [14]. Aerobic granules have been investigated as biosorbents for the removal of heavy metals, because the granules have high physical strength, high specific gravity, and good settleability [15,16]. However, most studies so far have been conducted using metal cations, based on removal resulting from electrostatic attraction between negatively charged granules and positively charged cations. This mechanism is less efficient if the ions to be removed bear negative charges, such as Sb(OH)6 − . A possible solution to this problem is to perform surface modification of the granules to change the nature of the surface charge and improve the removal of oxyanionic heavy metals. To date, only a few studies on the adsorption of oxyanions on modified aerobic granules have been carried out [17,18]. Because of cost and secondary pollution considerations, modification using common non-toxic inorganic salts is of particular interest; for example, a preliminary study of adsorption of Sb(V) on Fe-modified granules has been reported [19], in which the effects of pH, adsorption kinetics, and characteristics of the granules were explored. However, other important factors, including biomass dosage, shaker speed, ionic strength, and interfering ions, which could affect the performances of aerobic granules in Sb(V) adsorption, have not yet been fully investigated. Modification with FeCl3 changes the surface properties of the granules, and makes adsorption of Sb(OH)6 − on Fe-modified aerobic granules complicated, which has not been investigated in sufficient detail. These aspects are specifically addressed in this study. The objectives were (I) identification of factors affecting the adsorption capacity of granules; (II) further study of the adsorption isotherms and thermodynamics, to determine the adsorption characteristics of the granules; and (III) investigation of the adsorption mechanism of the negatively charged Sb(OH)6 − on Fe-modified aerobic granules. In this paper, we describe the macroscopic adsorption characteristics and microscopic removal mechanism. 2. Material and methods 2.1. Preparation and modification of aerobic granules Aerobic granules were cultivated in a column-type SBR, using a previously reported procedure [20]. The modification process was performed by mixing fresh granules with FeCl3 solution, as described in detail elsewhere [19]. Fe-modified granules were stored in tap water at 4 ◦ C for future use.
effect of ionic strength on the adsorption capacity. In addition, the interference effects of Cu2+ , Ni2+ , and Zn2+ at concentrations of 20, 40, and 60 mg/L on Sb(V) adsorption by granules were investigated. The test time was 300 min, and samples were collected at fixed time intervals for analysis. Each sample was centrifuged at 9000 × g for 5 min, and the Sb concentration in the supernatant was determined. The specific adsorption quantity, qt (mg Sb/g dry granules) and removal rate, r (% removal), are expressed as follows: qt = (C0 − Ce )v/m and r = (C0 − Ce )/C0 , where C0 and Ce are the initial and equilibrium concentrations, respectively, of Sb (mg/L), v is the volume of the solution (L), and m is the dry weight of modified granules (g). All experiments in this study were performed in triplicate. Metal-free and granules-free sample have been included as blank control. The results showed that the Sb(V) adsorbed onto the flasks was negligible, so was the Sb-release by granules. 2.4. Analytical methods The Sb(V) concentration was determined using inductively coupled plasma atomic emission spectroscopy (P-4010, Japan). The concentrations of Cu2+ , Ni2+ , and Zn2+ were determined using flame atomic absorption spectroscopy (Z-5000, Japan). The mass fraction of Fe in the granules was determined by dissolution and ICP-AES. 3. Results and discussion 3.1. Adsorption studies 3.1.1. Effect of biomass dosage The effect of biomass dosage on the adsorption of Sb(V) by Femodified aerobic granules is shown in Fig. 1. The Sb(V) removal rate increased with increasing biomass concentration from 2 to 30 g/L (wet basis); it reached a maximum of 99.3% at 20 g/L, and then remained almost unchanged. An increase in the biomass concentration results in a large surface area and more adsorption sites, which could bind more Sb(V) and improve the removal rate. However, the adsorption quantity decreased from 52.4 to 15.1 mg/g when the biomass concentration increased from 2 to 30 g/L. Liu et al. found that the ratio of initial Zn(II) concentration to initial granule concentration was closely associated with the biosorption quantity, and the ratio would better reflect the real driving force for metal biosorption by micro-organisms [21]. In present study, increasing biomass concentration at fixed initial Sb(V) concentration reduced the ratio, and accordingly weakened the driving force
2.2. Chemicals A standard stock solution of Sb(V) (1000 mg/L) was prepared by dissolving KSb(OH)6 in deionized water, and stock solutions of Cu2+ , Ni2+ , and Zn2+ (1000 mg/L) were prepared by dissolving CuSO4 , NiSO4 , and Zn(CH3 OOH)2 , respectively, in deionized water. The stock solutions were diluted to the required concentrations for the experiments, and the solution pHs were adjusted by adding 0.1 M HCl or NaOH. 2.3. Adsorption experiments The effects on adsorption of four major factors were studied. Different concentrations of modified granules, from 2 to 30 g/L (wet basis), were placed in flasks on a rotary shaker, and the shaking speed was varied from 0 to 175 rpm to investigate the effect of mass transfer in the bulk liquid. NaCl concentrations of 0.1 M and 0.3 M were supplied as background electrolytes, to evaluate the
Fig. 1. Effect of biomass dosage on the adsorption of Sb(V) on Fe-modified granules (T: 35 ◦ C, shaking speed: 175 rpm, C0 : 20 mg/L, pH: 3.4, contact time: 300 min).
L. Wang et al. / Journal of Hazardous Materials 284 (2015) 43–49
Fig. 2. Effect of shaking speed on the adsorption of Sb(V) on Fe-modified granules (T: 35 ◦ C, biomass dosage: 0.4 g/L, C0 : 20 mg/L, pH: 3.4).
for Sb(V) diffusion into the inside of the granules, which could lead to increasing unsaturation of adsorption sites per unit of biomass. When all Sb(V) was adsorbed on the granules, the contribution of additional biomass would be insignificant [22], and the specific adsorption quantity was therefore negatively correlated with the biomass dosage. These results suggest that the use of excessive amounts of biomass is uneconomic; a biomass dosage of 20 g/L (wet basis) gave the maximum efficiency in this study. 3.1.2. Effect of shaking speed External mixing of bulk liquid was expressed by shaking speed. Increasing the shaking speed increases the fluid shear stress, which may influence the boundary conditions and adsorption efficiency on the granule surface. Various shaking speeds were used to investigate the influence of shear stress, and the changes in adsorption quantity with time are shown in Fig. 2. The difference between adsorption efficiency at shaking speeds of 0 and 50 rpm was insignificant, because electrostatic interactions and film diffusion control the process under low shear, and a longer time was needed to reach adsorption equilibrium. When the shaking speed was increased to 100 rpm, the adsorption capacity of the Fe-modified granules improved significantly, and the adsorption quantity rose by 20% at 5 h. The increasing turbulence in the bulk liquid significantly improved mixing of the granules with soluble Sb(V), and the stronger driving force facilitated diffusion of Sb into the interiors of the granules. However, the use of a high speed, i.e., 175 rpm, did not improve the adsorption efficiency significantly, but a shorter time, i.e., 2 h, was needed to reach the adsorption equilibrium, compared with other speeds. In this study, a shaking speed of 175 rpm was therefore, used. 3.1.3. Effect of ionic strength NaCl was used to simulate various ionic strengths in the aqueous media. The adsorption quantity of Sb(V) on Fe-modified granules with time at different NaCl concentrations are shown in Fig. 3. A higher ionic strength significantly decreased the adsorption capacity, by 25% and 45% at 5 h with NaCl concentrations of 0.1 M and 0.3 M, respectively. There are several causes of this phenomenon. One reason is that at high concentrations, Cl− anion compete with Sb(OH)6 − for the available sites on the granule surface, and prevent Sb(V) approaching the granules. Second, Sb(V) exists mainly as highly polymerised hydroxy–chloro complexes or as colloidal hydrous oxides in the presence of high Cl− concentrations [13], so the original Sb(V) species may change. In addition, KSb(OH)6 is a
45
Fig. 3. Effect of ionic strength on the adsorption of Sb(V) on Fe-modified granules (T: 35 ◦ C, biomass dosage: 0.4 g/L, C0 : 20 mg/L, pH: 3.4, speed: 175 rpm).
common reagent for detecting Na+ in aqueous media, as follows [23]: NaCl + KSb(OH)6 → NaSb(OH)6 ↓ + KCl
(1)
The formation of a precipitate from these two reagents also changes the ionic state of Sb(V). The reasons stated above could individually, or in combination, lower the adsorption capacity of Fe-modified aerobic granules for Sb(V) on addition of NaCl as a background electrolyte. 3.1.4. Effect of interfering ions In the mining and exploitation of Sb, a large number of accompanying toxic cations, e.g., Cu2+ , Ni2+ , and Zn2+ , are commonly included in the Sb-containing wastewater stream produced [24]. Various concentrations of these cations were mixed with Sb(V) (20 mg/L) at pH 3.4 for 5 h. The adsorption quantity as a function of cation concentrations are shown in Fig. 4. The addition of interfering ions clearly inhibited adsorption of Sb(V) on the Fe-modified granules. The concentrations of Cu2+ , Zn2+ , and Ni2+ before and after adsorption were determined to investigate the cation-binding capacity of the Fe-modified granules. The results indicate that the concentrations of all the cations were almost unchanged after mixing with the granules.
Fig. 4. Effect of interfering ions on the adsorption of Sb(V) on Fe-modified granules (T: 35 ◦ C, biomass dosage: 0.4 g/L, C0 : 20 mg/L, pH: 3.4, speed: 175 rpm).
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Modification with FeCl3 made the surface of the aerobic granules positively charged, and prevented the cations from approaching the adsorbent. Cu2+ , Zn2+ , and Ni2+ , therefore, did not compete with Sb(OH)6 − for adsorption sites. Consequently, The inhibitory effect might be attributed to the corresponding anions by competing for adsorption sites with Sb(OH)6 − . Xi et al. found that the Sb(V) adsorption on bentonite was significantly affected in the presence of three competitive anions, and the effect was more evident in the case of PO4 3− and SO4 2− as compared to NO3 − [25]. The possible accumulation or precipitation of competitive anions on the surface of adsorbent could increase the negative surface charge and electrostatic repulsion with Sb(V), which might be the main reason of the inhibitory effect. In this study, CuSO4 , NiSO4 , and Zn(CH3 OOH)2 were used to generate interfering cations, therefore, the corresponding anions were SO4 2− , SO4 2− , and CH3 COO− , at molar concentrations of 0.34, 0.31, and 0.61 mmol/L, respectively, when the mass concentration of cations was all fixed at 20 mg/L. SO4 2− is an inorganic ion, while CH3 COO− is an organic weak acid anion, which could form CH3 COOH at acidic condition. The possible surface complexation between organic weak acid and Sb(V) could change the original species of Sb(OH)6 − , and reduce the adsorption capacity of the granules for Sb(V) [26]. Generally, the existence of anions could contribute to the decrease of Sb(V) adsorption onto the Fe-modified granules.
Table 1 Isotherm parameters for the adsorption of Sb(V) on Fe-modified granules. Langmuir
qm (mg/g) 125
b (L/mg) 0.056
R2 0.987
Freundlich
kF (mg/g) 27.1
n 3.9
R2 0.963
Temkin
B 12.2
A (L/mg) 18.2
R2 0.920
3.2. Isotherm analysis The adsorption of Sb(V) on Fe-modified granules was performed until equilibrium was reached, and the data were analyzed using three isotherm models, i.e., the Langmuir, Freundlich, and Temkin models. The Langmuir isotherm model assumes that the adsorbent surface has homogeneous sites, and monolayer adsorption occurs without interactions between adsorbed molecules [27]. The linear form of the Langmuir isotherm model can be expressed as [28]: Ce 1 Ce = + qm qe (bqm )
(2)
where Ce is the equilibrium concentration of Sb(V) (mg/L), qe is the equilibrium adsorption quantity (mg/g), qm is the maximum adsorption quantity (mg/g), and b is the Langmuir equilibrium constant (L/mg). The Freundlich isotherm model is used to describe adsorption on heterogeneous surface through multilayer adsorption [20], which can be described as [29]: lnqe =
1 n ln C e + lnkF
(3)
where kF is the Freundlich constant, representing the adsorption capacity (mg/g), and 1/n is related to the adsorption intensity of the adsorbent. The Temkin isotherm model assumes that the heat of adsorption of the molecules decreases linearly with coverage, because of the interactions between the adsorbent and adsorbate, and a uniform distribution of binding energies characterizes the adsorption up to a maximum energy; this model can be expressed as [30]: qe = BlnC e +BlnA
(4)
where B is related to the heat of adsorption and A represents the equilibrium binding constant (L/mg), corresponding to the maximum binding energy. Adsorption isotherms are indispensable in investigating the nature of the interactions between adsorbents and adsorbates [22]. The equilibrium data were fitted to the Langmuir, Freundlich, and Temkin isotherm models; the results are shown in Table 1 and Fig. 5. The order of the correlation coefficients, R2 , for the biosorption
Fig. 5. Adsorption isotherm of Sb(V) on Fe-modified granules by different models.
of Sb(V) is Langmuir > Freundlich > Temkin, therefore, the experimental equilibrium adsorption data are best fitted by the Langmuir model. The maximum adsorption quantity and Langmuir constant were calculated to be 125 mg/g and 0.056 L/mg, respectively. This indicated that adsorption of Sb(V) on the Fe-modified granules mainly occurred in a monolayer, without interactions between adsorbed molecules; these results are similar to those reported for biosorption of Cu2+ on modified Penicillium biomass [31]. However, based on the previous studies, the removal rate of Sb(V) in aqueous solution by original and Fe-modified granules were 14% and 99% at pH 3.4, respectively [19]. This suggested that the calculated maximum adsorption quantity of 125 mg/g in the present study could be the result of combined action by original granules and Fe-modification. 3.3. Thermodynamics analysis In thermodynamics, it is assumed that energy cannot be gained or lost in an isolated system, and the entropy change represents the driving force [32]. The changes in entropy (S) and enthalpy (H) are important thermodynamics parameters for the identification of spontaneous process [33]. The thermodynamic parameters S, H, and the Gibbs free energy (G) can be calculated using the following equations [27]: G = −RT lnK K=
qe Ce
G = H − TS
(5) (6) (7)
where R is the ideal gas constant [8.314 × 10−3 kJ/(mol K)], T is the absolute temperature (K), and K is the distribution coefficient, which is the ratio of the equilibrium adsorption quantity to the equilibrium concentration of Sb(V).
L. Wang et al. / Journal of Hazardous Materials 284 (2015) 43–49
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Table 2 Thermodynamic parameters for the adsorption of Sb(V) on Fe-modified granules. T (K)
G kJ/(mol K)
H kJ/(mol K)
S kJ/(mol K)
283 298 308 318
−4.87 −5.98 −6.72 −7.46
16.1
74.1
chemisorption [33]. The positive value of S suggests an increase in randomness at the solid/liquid interface during adsorption of Sb(V) on the Fe-modified granules. The values of G, H, and S are shown in Table 2. 3.4. Test of acid-soluble Fe in Fe-modified granules
Fig. 6. Plots of lnK versus 1/T for adsorption of Sb(V) on Fe-modified granules.
The final equation can be written as: lnK =
S H − R (RT )
(8)
The entropy change (S) and enthalpy change (H) can be calculated from the slope and intercept of a plot of 1/T versus lnK, as shown in Fig. 6. The Gibbs free energy (G) can be determined from H, S, and T using Eq. (7). The negative values of G indicate that adsorption of Sb(V) on Fe-modified granules is spontaneous in the temperature range of 283–318 K. The absolute value of G increases with increasing temperature, which suggests that adsorption occurs easily at high temperatures. The positive value of H indicates that Sb(V) adsorption on Fe-modified granules is endothermic, which agrees with the calculated results for the Gibbs free energy. Ions that dissolve well in water need to be dehydrated to be adsorbed on an adsorbent. When the dehydration process requires energy exceeding the exothermicity of ion attachment to the surface, a high temperature provides more energy for dehydration of Sb(V) and more activation energy for monolayer
Various initial concentrations of Sb(V) were used in the adsorption experiments. After adsorption, the Sb-loaded and the original granules were firstly dissolved in nitrohydrochloric acid, and precipitations were found in all samples except the granules without Sb dosage. The pretreatment was followed by drying, carbonization, and ashing in muffle furnace. The residues were further immersed in nitrohydrochloric acid, and the sample unloaded with Sb could be dissolved completely, leaving clear acid solution. However, insoluble precipitation could be observed in all the other samples. The concentrations of Fe in the supernatant were further determined using ICP-AES. The mass fractions of acid-soluble Fe in the Sbbearing granules were calculated and shown in Fig. 7. Before adsorption, the concentration of Fe in the modified granules was 5.9% (w/w). With increasing initial Sb(V) concentration, the concentration of acid-soluble Fe in the granules decreased. For initial concentrations of 300 and 500 mg/L, the amounts of acidsoluble Fe decreased to 2.4% and 2.3%. This could be attributed to the possible formation of stable complexes, which cannot be easily broken up by acid. In addition, the higher the initial concentration of Sb(V) is, the more Fe on the granules could be participating in adsorption, hence the lower Fe concentration in the supernatant. However, some Fe in the inner layers of the granules might not be
Fig. 7. Concentration of acid-soluble iron in Fe-modified granules.
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accessible to Sb(V), which might be why there was still a certain level of acid-soluble Fe under high Sb(V) concentration. 3.5. Adsorption mechanism
protonation, and Fe OH is positively charged, whereas deprotonation occurs at high pH, and Fe OH is negatively charged. The reaction equations can be expressed as: Fe OH + H+ →
It has been reported that the macroscopic evidence of formation of inner-sphere complexes is the adsorption process’ insensitivity to changes of ionic strength; otherwise, it was more possible that out-sphere complexes was formed [34]. However, Xi et al. studied the adsorption of Sb(V) on kaolinite, and found that different ionic strengths had different effects on adsorption, which suggested that both inner-sphere and outer-sphere complexes were formed [34]. Similarly the influence of ionic strength in this study was found to be complicated, therefore the adsorption mechanism of Sb(V) on the Fe-modified granules could also be multiple and complicated. In a previous study, a similar modification of biomass was performed using FeCl3 , and it was found that Fe(III) existed in the form of Fe OH on biomass surface [35,36]. Therefore, it could be inferred that in present work, hydrolysis of Fe(III) in the solution could lead to formation of Fe(OH)3 , which were then stably attached onto the granules. Fe-modified granules thus have a large amount of Fe-hydroxyl groups, i.e., Fe OH, on the surface. The speciation of Fe OH depends on solution pH. A low pH leads to
−
Fe OH + OH →
Fe OH2 + −
Fe O + H2 O
(9) (10)
On the other hand, some researchers suggested that adsorption of Sb(V) onto hydrous ferric oxide (HFO) can be described by an inner-sphere surface complex model, the reaction of which is as follow [37]: Fe OH + Sb(OH)6 − = FeOSb(OH)5 − + H2 O
(11)
Similarly, the formation of complex between OH and Sb(OH)6 – by dehydration could contribute significantly to Sb(V) removal in present study. In theory, one Fe atom could have more than one OH group attached to it, and could bind more than one Sb atom, which was the reason why Fe-modified granules had a high adsorption capacity for Sb(V). The mechanism of adsorption of Sb(V) on Fe-modified granules therefore could be inferred, based on the results of our experiments and previous reports [35,36,38]. Under acidic condition, Fe-modified granules tend to be protonated to Fe OH2 + ·Sb(OH)6 − can therefore be
Fig. 8. The four steps of adsorption of Sb(V) on Fe-modified granules.
L. Wang et al. / Journal of Hazardous Materials 284 (2015) 43–49
easily attracted by positively charged granules through electrostatic interaction and approach the granules surface. Outer-sphere complexes [ Fe OH2 + ·Sb(OH)6 − ] can be formed through ion exchange, followed by formation of more stable chemical bonds ( Fe O Sb), and inner-sphere complexes are finally formed. The four possible steps in adsorption of Sb(V) on Fe-modified aerobic granules are shown in Fig. 8. 4. Conclusion Fe-modified aerobic granules were excellent biosorbents for the removal of Sb(V). Adsorption batch experiments suggested increasing the biomass dosage increased the number of adsorption sites and further enhanced the removal rate of Sb(V), but decreased the specific adsorption capacity. Increasing the shaking speed improved the adsorption capacity of the biosorbent and shortened the equilibrium adsorption time. The adsorption capacity of the Fe-modified granules was found to be inhibited significantly by addition of foreign ions, and the Fe-modified granules had almost no affinity for cations in aqueous media. The Langmuir model best fitted the equilibrium data, with a monolayer adsorption quantity of 125 mg/g. The adsorption of Sb(V) on the Femodified granules was found to be spontaneous and endothermic. Surface protonation, electrostatic interactions, ion exchange, and surface complexation were inferred to be the main mechanisms of adsorption of Sb(V) on Fe-modified granules. References [1] F. Sun, F. Wu, H. Liao, B. Xing, Biosorption of antimony(V) by freshwater cyanobacteria Microcystis biomass: chemical modification and biosorption mechanisms, Chem. Eng. J. 171 (2011) 1082–1090. [2] F. Wu, F. Sun, S. Wu, Y. Yan, B. Xing, Removal of antimony(III) from aqueous solution by freshwater cyanobacteria Microcystis biomass, Chem. Eng. J. 183 (2012) 172–179. [3] M. He, J. Yang, Effects of different forms of antimony on rice during the period of germination and growth and antimony concentration in rice tissue, Sci. Total Environ. 243 (1999) 149–155. [4] Z. Wu, M. He, X. Guo, R. Zhou, Removal of antimony(III) and antimony(V) from drinking water by ferric chloride coagulation: competing ion effect and the mechanism analysis, Sep. Purif. Technol. 76 (2010) 184–190. [5] W. Xu, H. Wang, R. Liu, X. Zhao, J. Qu, The mechanism of antimony(III) removal and its reactions on the surfaces of Fe–Mn binary oxide, J. Colloid Interface Sci. 363 (2011) 320–326. [6] X.J. Guo, Z.J. Wu, M.C. He, Removal of antimony(V) and antimony(III) from drinking water by coagulation–flocculation–sedimentation (C–F–S), Water Res. 43 (2009) 4327–4335. [7] A.K. Leuz, C.A. Johnson, Oxidation of Sb(III) to Sb(V) by O2 and H2 O2 in aqueous solutions, Geochim. Cosmochim. Acta 69 (2005) 1165–1172. [8] F. Quentel, M. Filella, C. Elleouet, C.L. Madec, Kinetic studies on Sb(III) oxidation by hydrogen peroxide in aqueous solution, Environ. Sci. Technol. 38 (2004) 2843–2848. [9] M.A. Salam, R.M. Mohamed, Removal of antimony(III) by multi-walled carbon nanotubes from model solution and environmental samples, Chem. Eng. Res. Des. 91 (2013) 1352–1360. [10] M. Iqbal, A. Saeed, R.G. Edyvean, Bioremoval of antimony(III) from contaminated water using several plant wastes: optimization of batch and dynamic flow conditions for sorption by green bean husk (Vigna radiata), Chem. Eng. J. 225 (2013) 192–201. [11] Y. Leng, W. Guo, S. Su, C. Yi, L. Xing, Removal of antimony(III) from aqueous solution by graphene as an adsorbent, Chem. Eng. J. 211 (2012) 406–411. [12] N. Belzile, Y.W. Chen, Z.J. Wang, Oxidation of antimony(III) by amorphous iron and manganese oxyhydroxides, Chem. Geo. 174 (2001) 379–387.
49
[13] M. Filella, N. Belzile, Y.W. Chen, Antimony in the environment: a review focused on natural waters: II. Relevant solution chemistry, Earth-Sci. Rev. 59 (2002) 265–285. [14] D.W. Gao, L. Liu, H. Liang, W.M. Wu, Comparison of four enhancement strategies for aerobic granulation in sequencing batch reactors, J. Hazard. Mater. 186 (2011) 320–327. [15] Y. Liu, H. Xu, Equilibrium, thermodynamics, and mechanisms of Ni2+ biosorption by aerobic granules, Biochem. Eng. J. 35 (2007) 174–182. [16] L. Yao, Z.F. Ye, M.P. Tong, P. Lai, J.R. Ni, Removal of Cr3+ from aqueous solution by biosorption with aerobic granules, J. Hazard. Mater. 165 (2009) 250–255. [17] X.F. Sun, C. Liu, Y. Ma, S.G. Wang, B.Y. Gao, X.M. Li, Enhanced Cu(II) and Cr(VI) biosorption capacity on poly (ethylenimine) grafted aerobic granular sludge, Colloid Surf. B 82 (2011) 456–462. [18] X.F. Sun, Y. Ma, X.W. Liu, S.G. Wang, B.Y. Gao, X.M. Li, Sorption and detoxification of chromium(VI) by aerobic granules functionalized with polyethylenimine, Water Res. 44 (2010) 2517–2524. [19] L. Wang, C. Wan, D.J. Lee, X. Liu, Y. Zhang, X. Chen, J.H. Tay, Biosorption of antimony(V) onto Fe(III)-treated aerobic granules, Bioresour. Technol. 158 (2014) 351–354. [20] L. Wang, C. Wan, D.J. Lee, J.H. Tay, X. Chen, X. Liu, Y. Zhang, Adsorption–desorption of strontium from waters using aerobic granules, J. Taiwan Inst. Chem. Eng. 44 (2013) 454–457. [21] Y. Liu, S.F. Yang, S.F. Tan, Y.M. Lin, J.H. Tay, Aerobic granules: a novel zinc biosorbent, Lett. Appl. Microbiol. 35 (2002) 548–551. [22] F. Sun, W.L. Sun, H.M. Sun, J.R. Ni, Biosorption behavior and mechanism of beryllium from aqueous solution by aerobic granule, Chem. Eng. J. 172 (2011) 783–791. [23] X.Z. Cao, M.L. Zhao, D.W. Wang, The influences of EMP irradiation of rat hypothalamus in vitro on intracellular LDH and LDH, AST, CHE, K+ , and Na+ in culture medium, Acta Biophy. Sin. 18 (2002) 302–306. [24] X.Y. Shan, S.Q. Xu, The analysis and assessment on the pollution condition of heavy metals in the soil around the Qinglong Dachang antimony mining area in Guizhou province, J. Guizhou Univ. (Nat. Sci.) 28 (2011) 132–135. [25] J.H. Xi, M.C. He, C.Y. Lin, Adsorption of antimony(III) and antimony(V) on bentonite: kinetics, thermodynamics and anion competition, Microchem. J. 97 (2011) 85–91. [26] V.K. Mittal, S. Bera, S. Narasimhan, S. Velmurugan, Adsorption behavior of antimony(III) oxyanions on magnetite surface in aqueous organic acid environment, Appl. Surf. Sci. 266 (2013) 272–279. [27] D. Ozdes, C. Duran, H.B. Senturk, Adsorptive removal of Cd(II) and Pb(II) ions from aqueous solutions by using Turkish illitic clay, J. Environ. Manage. 92 (2011) 3082–3090. [28] I. Langmuir, The adsorption of gases on plane surfaces of glass, mica, and platinum, J. Am. Chem. Soc. 40 (1918) 1361–1403. [29] H. Freundlich, Über Die Adsorption in Lösungen, Engelmann, Leipzig, 1906. [30] M. Temkin, V. Pyzhev, Kinetics of ammonia synthesis on promoted iron catalysts, Acta Physiochim. U.R.S.S. 12 (1940) 217–222. [31] X. Zhang, H.J. Su, T.W. Tan, G. Xiao, Study of thermodynamics and dynamics of removing Cu(II) by biosorption membrane of Penicillium biomass, J. Hazard. Mater. 193 (2011) 1–9. [32] K.V. Kumar, A. Kumaran, Removal of methylene blue by mango seed kernel powder, Biochem. Eng. J. 27 (2005) 83–93. [33] M.V. Subbaiah, G. Yuvaraja, Y. Vijaya, A. Krishnaiah, Equilibrium, kinetic, and thermodynamic studies on biosorption of Pb(II) and Cd(II) from aqueous solution by fungus (Trametes versicolor) biomass, J. Taiwan Inst. Chem. Eng. 42 (2011) 965–971. [34] J. Xi, M. He, C. Lin, Adsorption of antimony(V) on kaolinite as a function of pH, ionic strength and humic acid, Environ. Earth Sci. 60 (2010) 715–722. [35] M. Aryal, M. Ziagova, M. Liakopoulou-Kyriakides, Study on arsenic biosorption using Fe(III)-treated biomass of Staphylococcus xylosus, Chem. Eng. J. 162 (2010) 178–185. [36] M. Aryal, M. Ziagova, M. Liakopoulou-Kyriakides, Comparison of Cr(VI) and As(V) removal in single and binary mixtures with Fe(III)-treated Staphylococcus xylosus biomass: thermodynamic studies, Chem. Eng. J. 169 (2011) 100–106. [37] X.J. Guo, Z.J. Wu, M.C. He, Removal of antimony(V) and antimony(III) from drinking water by coagulation–flocculation–sedimentation (C–F–S), Water Res. 43 (2009) 4327–4335. [38] A.K. Leuz, H. Mönch, C.A. Johnson, Sorption of Sb(III) and Sb(V) to goethite: influence on Sb(III) oxidation and mobilization, Environ. Sci. Technol. 40 (2006) 7277–7282.