Removal of Aqueous Hg(II) by Polyaniline: Sorption Characteristics


Jun 9, 2009 - ... P. R. China, Department of Civil and Environmental Engineering, ... Citation data is made available by participants in CrossRef's Ci...
0 downloads 0 Views 309KB Size


Environ. Sci. Technol. 2009, 43, 5223–5228

Removal of Aqueous Hg(II) by Polyaniline: Sorption Characteristics and Mechanisms J I N G W A N G , †,§ B A O L I N D E N G , ‡ HUAN CHEN,† XIAORONG WANG,† AND J I A N Z H O N G Z H E N G * ,† State Key Laboratory of Pollution Control and Resource Reuse, School of the Environment, Nanjing University, Nanjing, 210093, P. R. China, Department of Civil and Environmental Engineering, University of Missouri, Columbia, Missouri 65211, and Institute of Chemistry, Henan Academy of Sciences, Zhengzhou, 450002, P. R. China

Received December 30, 2008. Revised manuscript received April 23, 2009. Accepted May 18, 2009.

A polyaniline (PAN) prepared by chemical oxidation method was studied for Hg(II) removal from aqueous solutions. Batch adsorption results showed solution pH values had a major impact on mercury adsorption by this sorbent with optimal removal observed around pH 4-6. At both acidic and alkaline solutions beyond this optimal pH window, sorption capacity ofPANwassubstantiallylowered,withtheimpactlesspronounced at pH above 6. Among the water constituents tested, only chloride and humic acid had significant inhibition on mercury removal due to competitive complexation. In the range of 0.02-0.2 M, ionic strength had less impact on Hg(II) removal by PAN while further increase in background electrolyte concentration to 1.0 M substantially decreased mercury removal. An adsorption mechanism was proposed by analyzing the XPS spectra of the key elements (N1s, Cl2p and Hg4f) on polyaniline surfaces and the change of its electrokinetic properties, both before and after Hg(II) adsorption. Specifically, at pH 5.5, it is likely that all the nitrogen-containing functional groups on the polymer matrix including imine, protonated imine and amine could be responsible for mercury adsorption, with imine having the highest affinity while the remaining two having similar strength to complex mercury.

Introduction Mercury, one of the most toxic metals, has drawn much attention due to its toxicity and impact on the public health. The main anthropogenic sources contributing to mercury contamination include wastewater discharges and atmospheric deposition from mining activities, oil and coal combustion, chlor-alkali industries, cement production, municipal waste and sewage sludge combustion, and manufacture of batteries, among others (1-4). Mercury released to the environment can be further transformed through both biotic and abiotic methylation to methylmercury (5), a substance much more toxic than inorganic mercury. Methylmercury can also undergo demethylation through either * Corresponding author phone/fax: 086-25-83594492; e-mail: [email protected] † Nanjing University. ‡ University of Missouri. § Henan Academy of Sciences. 10.1021/es803710k CCC: $40.75

Published on Web 06/09/2009

 2009 American Chemical Society

bacterial or abiotic photochemical pathways (6). Because of its high toxicity, mercury has been included on the list of priority pollutants by U.S. EPA (7) with a mandatory discharge limit of 10 µg/L for wastewater (2) and a maximum concentration level of 2 µg/L for drinking water (8). Conventional technologies for treatment of aqueous mercury include precipitation, coagulation, reduction, membrane separation, ion exchange, and adsorption (2, 4). Among these technologies, adsorption has been widely studied because it is easy to operate and cost-effective. Many adsorbents have been studied for Hg(II) removal from aqueous solutions, including activated carbons (9, 10), silicates (11-13), polymers (14), and biomass, such as bacteria (15) and chitosan (16, 17). In addition, other materials including clays, organic matters, iron oxides, and pyrite, belonging to the general categories of soils (18), suspended solids/sediments in natural waters (19-21), and natural minerals (22, 23) have also been extensively studied for their adsorption properties for mercury. Materials effective for mercury removal generally carry sulfur, nitrogen and oxygencontaining functional groups as major binding sites for mercury. For example, strong interactions between mercury and organic matters have been attributed to the binding of mercury with reduced sulfur-containing functional groups in these substances (24, 25). Polyaniline, a well-known conducting polymer, has been subjected to extensive studies in lightweight battery electrodes, sensors, electromagnetic shielding devices, and anticorrosion coatings (26) because of its controllable electrical conductivity, environmental stability, good redox reversibility, and low cost. Depending on the ratio of reduced benzenoid amine to oxidized quinoid imine, polyaniline can exist as fully reduced leucoemeraldine, fully oxidized pernigraniline, and half oxidized emeraldine. Moreover, emeraldine base can be reversibly doped preferentially on the imine nitrogen atoms with a protonic acid to yield the conductive emeraldine salt (26). Because it carries large amounts of amine and imine functional groups, this polymer is expected to have interactions with some metal ions having strong affinity to nitrogen. For instance, polyaniline has been effectively used to preconcentrate and separate metals before instrumental analysis (27, 28). The redox reaction between polyaniline and Hg(I) has been exploited to prepare a metallic-polymer composite for trace metal analysis by anodic stripping voltammetry (29). However, compared with other areas of inquiry, this material is much less understood in terms of environmental applications. Among the few studies reported, polyaniline and its composites have been explored for removal of mercury from aqueous solutions (30-34). In one study, Lu et al. (34) used copolymer aniline-sulfophenylenediamine for Hg(II) adsorption from aqueous solutions, achieving an adsorption capacity of 497.7 mg/g. It is generally accepted in the literature that nitrogen atoms on polyaniline surfaces are responsible for Hg(II) adsorption. However, systematic studies on sorption properties of Hg(II) on polyaniline are still lacking, and the associated adsorption mechanisms are yet to be clarified using solid spectroscopic methods. Knowledge on these aspects would potentially facilitate developing novel technologies for effective contaminant removal from mercurycontaining waters. The objectives of the current study were to (1) characterize the sorption properties of polyaniline prepared by chemical oxidation for Hg(II) under a range of water chemistry conditions and (2) investigate the associated sorption mechanisms. A series of batch sorption tests were conducted to VOL. 43, NO. 14, 2009 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

9

5223

examine the effects of pH, ionic strength, and water constituents on Hg(II) uptake by this macromolecule. The associated adsorption mechanisms were proposed by analyzing the experimental data from XPS spectra of the key elements involved and characterizing the electrokinetic properties of polyaniline, both before and after Hg(II) adsorption.

Experimental Section Materials. Mercury nitrate monohydrate (Hg(NO3)2 · H2O) and sodium humate (designated NaHA) of ACS reagent grade were purchased from Sigma-Aldrich. All other reagents of analytical grade were purchased from Sinopharm Chemical Reagent Co., Ltd., Shanghai, China. All solutions were prepared with high-purity water (18.2 MΩ · cm). NaNO3 was used to adjust the ionic strength, and HNO3 and NaOH were used to adjust the solution pH throughout the experiments. Polyaniline was prepared by the chemical oxidation method following the procedure reported by Shimano and MacDiarmid (26). Compared with electrochemical polymerization, this method is preferred for large-scale synthesis because it is easy for scale-up (35) and has a high yield of polyaniline (30). Briefly, 0.11 mol of aniline was dissolved in 300 mL of 1 M HCl to which 0.11 mol of ammonium peroxodisulphate was added under continuous stirring. The insoluble dark-green precipitate thus formed was washed with copious deionized water until water became colorless. The resultant material was kept in a desiccator prior to use (designated PAN). FTIR spectrum (Figure S1, Supporting Information) suggested the synthesized PAN was in its emeraldine salt state (36). Apparently, the counteranion of this PAN was chlorine. The specific surface area of this PAN was 35.4 m2/g. Batch Sorption Experiments. A preliminary kinetic test demonstrated that adsorption equilibrium of mercury on PAN was achieved within approximately 24 h under the experimental conditions. Therefore, an equilibrium time of 24 h was adopted for all the adsorption experiments. Batch adsorption experiments were conducted to investigate Hg(II) adsorption as a function of initial mercury concentration (10-250 mg/L), aqueous pH (1-11), ionic strength (0.02, 0.2, 1.0 M NaNO3), and water constituents (chloride, sulfate, phosphate, calcium, magnesium, and humic acid), all in duplicate. Except for humic acid (1-100 mg/L), all other constituents were investigated in concentration ranges of 0.1-30 mM. To start the batch adsorption experiment, 10 mg of PAN were mixed with 50 mL of aqueous Hg(II) solution contained in a plastic centrifuge tube, which was secured on a shaking table working at 25 °C and 150 rpm. After reaching equilibrium, the mixture was immediately filtered with a 0.45 µm membrane, and the filtrate was collected, acidified in 5 vol % of HNO3, and analyzed within 24 h for mercury concentration by atomic fluorescence spectrophotometry (AF-600, Beijing Rayleigh Analytical Instrument Corp. Beijing, China). Every sample was analyzed in duplicate. Adsorption capacity (Qe, mg/g) and removal rate (R, %) for mercury by PAN at equilibrium were calculated by conducting a mass balance of mercury before and after adsorption. Sample Characterization. The electrokinetic mobility of PAN particles (0.01 M NaNO3, pH 4-11) was measured by ZetaPALS (Brookhaven Instruments, Holtsville, NY). Solution pH of the PAN suspension was adjusted with 0.01-0.1 M HNO3 and NaOH solutions. XPS measurements of polyaniline before (designated PAN) and after (PAN-Hg) Hg(II) sorption were performed on a VG ESCALB MK-II Instrument (VG, UK). Both survey and high-resolution spectra of N1s, Cl2p, and Hg4f were collected and calibrated to the binding energy (BE) of C1s at 284.6 eV. The calibrated high-resolution spectra were fitted by XPSpeak 4.1 software (37) and surface elemental stoichiometries were determined from the peak-area ratios 5224

9

ENVIRONMENTAL SCIENCE & TECHNOLOGY / VOL. 43, NO. 14, 2009

FIGURE 1. Sorption isotherms of Hg(II) by PAN at three levels of ionic strength. 0.02 M (9), 0.2 M (O), and 1.0 M (2) NaNO3; Hg(II) initial concentrations (C0) ) 10-250 mg/L; Ce for concentration of Hg(II) at equilibrium; pH ) 5.5 buffered by 2 mM NaAc/HAc solution; PAN dosage ) 0.2 g/L.

FIGURE 2. Impact of solution pH on mercury removal by PAN. C0 of Hg(II) ) 43.0 mg/L; ionic strength ) 0.2 M NaNO3; PAN dosage ) 0.2 g/L; pH adjusted by HNO3 and NaOH solutions; pHe for solution pH at equilibrium. corrected by sensitivity factor. A blank test (designated PAN 5.5) was also conducted under the same experimental conditions (i.e., pH ) 5.5, ionic strength ) 0.2 M NaNO3) as for preparation of PAN-Hg, but without Hg(II) addition. Additionally, desorption of Hg(II) from the mercury sorbed polyaniline (PAN-Hg-1.0 M HCl) with 1.0 M HCl solution was studied to better understand the affinity of mercury to different binding sites on polyaniline.

Results Adsorption Isotherms. Adsorption of Hg(II) by PAN was studied at pH 5.5 covering a wide range of ionic strengths (0.02, 0.2, and 1.0 M NaNO3) and Hg(II) concentrations. As illustrated in Figure 1, the two sorption isotherms at 0.02 and 0.2 M ionic strengths had no obvious difference. The maximum uptake capacities for mercury (Qmax) at both ionic strengths were about 600 mg/g, a little higher than that reported (497.7 mg/g) for the copolymer of aniline and sulfophenylenediamine (34). When ionic strength was further increased to 1.0 M, Qmax of PAN was significantly reduced to about 400 mg/g. Effect of pH. Solution pH had a significant impact on mercury removal, with apparent inhibition observed in both alkaline and very acidic solutions (Figure 2). For instance, the mercury removal ratio increased abruptly from less than 20% at pH 2.0 to about 90% at pH 4.0. Increasing the solution pH above 6.0 also decreased the adsorption, although the extent was less pronounced. This differed from the study by Balarama Krishna et al. (32), in which they found that mercury

FIGURE 3. Impact of water constituents on Hg(II) removal by PAN. C0 of Hg(II) ) 47.6 mg/L, ionic strength ) 0.2 M NaNO3; pH ) 5.5 buffered by 2 mM NaAc/HAc solution; PAN dosage ) 0.2 g/L. adsorption by polyaniline was independent of solution pH, probably due to the lower mercury concentration they used. Observations of inhibition on Hg(II) sorption have been well documented for both acidic (15, 18, 38) and basic solution systems (15, 18). Effect of Water Constituents. As illustrated in Figure 3, both Cl- and humic acid in aqueous solutions inhibited Hg(II) adsorption, with the effect of Cl- being more pronounced. At initial Hg(II) concentration of 47.6 mg/L, 20 mM of Cldecreased the removal ratio to less than 20%, while much less inhibition (to 70%) was observed in the presence of 20 mg/L of NaHA. In comparison, the effects of sulfate, phosphate, calcium and magnesium were insignificant even at 30 mM concentration levels. XPS Analysis. XPS spectra of both survey and highresolution scans for the key elements on polyaniline surfaces before and after adsorption were studied to gain the insights into the adsorption mechanism(s) of Hg(II) on this sorbent material. As shown in the survey spectra in Figure 4A, the presence of Hg4f in PAN-Hg clearly confirmed the adsorption of mercury after PAN was equilibrated with Hg(II) solution. The appearance of Hg4f was in accordance with the decrease in N1s peak area. After elution of the sorbed mercury from PAN-Hg with 1 M HCl, its corresponding survey spectra (PANHg-1.0 M HCl) showed an apparent decrease in Hg4f peak area and a corresponding increase in N1s peak area. The results suggested that sorption of Hg(II) was mainly with nitrogen, so desorption of Hg(II) freed nitrogen and led to an increase in N1s peak. It was also observed that chlorine could be partially dedoped from the polymer matrix as indicated by the decrease in the Cl2p peak area for PAN 5.5 as compared with PAN. The presence of Hg(II) in the solution, however, helped to retain most of the chlorine on the polymer matrix (Figure 4A-b). High-resolution spectra (Figure 4C) showed that N1s could be fitted into three major species with varied fractions under different experimental conditions. Peak fitting illustrated that N1s of PAN (Figure 4C-a) could be grouped into two peaks at 399.5 and >400.0 eV, corresponding to nitrogen atoms in amine (-NH-) and doped imine (-NH+ · -) functional groups, respectively (39). Nearly similar peak areas for these two nitrogen species suggested that this PAN was in its emeraldine salt state (26, 39). Surface elemental stoichiometries tabulated in Table 1 illustrated that total nitrogen contents decreased from 11.1 to 9.8 atom % after Hg(II) sorption on PAN. It is interesting to observe that exposure of PAN to pH 5.5 aqueous solution (Figure 4C-c) caused similar changes in the N1s spectra but having different distribution of nitrogen species (Table 1).

FIGURE 4. XPS spectra of survey scan (A) and high-resolution scan of Hg4f (B), N1s (C), and Cl2p (D) for polyaniline samples with different treatments: (a) PAN; (b) PAN-Hg, PAN equilibrated with Hg(II) solution (pH ) 5.5, ionic strength ) 0.2 M NaNO3, C0 of Hg(II) ) 100 mg/L); (c) PAN 5.5, blank test prepared by equilibrating PAN with aqueous solution under the same experimental conditions as PAN-Hg (pH 5.5, ionic strength 0.2 M NaNO3) but without Hg(II) addition; (d) PAN-Hg-1.0 M HCl, after eluting PAN-Hg for 24 h with 1.0 M HCl solution. Cl2p spectra were also subjected to substantial changes after different treatments of the polymer. Peak fitting (Figure 4D) demonstrated that Cl2p spectra could be further divided into 3 spin-orbit split doublets (Cl2p3/2 and Cl2p1/2) due to Cl-, Cl*, and C-Cl, which represent chloride anion, anionic chloride resulting from the charge transfer between chlorine and PAN polymer chain, and the covalently bonded chlorine, respectively (39). As illustrated, their Cl2p3/2 were situated at BEs of 197.3 ( 0.1, 198.3 ( 0.1, and 200.6 ( 0.2 eV, respectively (39). In agreement with deprotonation of imine nitrogen observed for PAN 5.5, Cl-, and Cl* species disappeared from Cl2p spectra for PAN 5.5, causing substantial decrease in total chlorine content from 3.6 to 0.9 atom % (Table 1). The residual chlorine was mainly in covalent-bond chlorine form. However, desorption of chloride from PAN could be inhibited by the presence of Hg(II). Hg4f spectra of PAN-Hg (Figure 4B) could be fitted with two doublet-peaks, split both by 4.0 eV (the distance between Hg4f7/2 and Hg4f5/2 peaks). Hg4f7/2 for the two peaks were situated at 101.1 (designated Hg1) and 103.0 eV (designated Hg2), respectively, each with a relatively broad fwhm (full width at half-maximum) of over 2.0 eV. Surface elemental stoichiometries calculated from the peak-area ratios showed that the atomic ratio of Hg1 to Hg2 was about 1.4 (1.7:1.2). After desorption with 1.0 M HCl solution (PAN-Hg-1.0 M HCl), contents of Hg1 and Hg2 were both decreased, with the decrease in Hg2 more pronounced (Table 1). Electrophoretic Mobility of Polyaniline. Electrophoretic mobilities of polyaniline both before (PAN) and after (PANHg) Hg(II) sorption were measured as a function of solution VOL. 43, NO. 14, 2009 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

9

5225

TABLE 1. Distribution of N, Cl, and Hg Species on Polyaniline Surfacesa

atom % sample

PAN PAN-pH 5.5 PAN-Hg

Cl2p3/2, BE, eV atom % sample

PAN PAN-pH 5.5 PAN-Hg

Hg4f7/2, BE, eV PAN-Hg PAN-Hg-1.0 M HCl decrease ratio of Hgi, % atom % sample a

1.4 2.9

-NH-, 399.5 5.1 5.9 4.5

-NH+ · -, > 400 6.0 3.7 2.4

total N 11.1 11.0 9.8

Cl-, 197.3 ( 0.1 1.8

Cl*, 198.3 ( 0.1 0.6

C-Cl, 200.6 ( 0.2 1.2 0.9 2.4

total Cl 3.6 0.9 3.6

Hg2, 102.2-103.0 1.2 0.5 58.3

total Hg 2.9 1.8

-Nd, 398.3

N1s, BE, eV

1.2 Hg1, 100.6-101.1 1.7 1.3 23.5

BE for binding energy.

pH. As shown in Figure 5A, PAN had a pHPZC about 5.8. After Hg(II) sorption, pHPZC of PAN-Hg shifted to a higher pH at 6.7, possibly due to specific adsorption of Hg(II) to the binding sites on PAN, changing its surface charge. The charging mechanism of PAN could be better understood with the aid of Figure 5B. At low pHs, nitrogen atoms, mainly on the imine groups, which are relatively easier to be protonated than the amine groups (39, 40), were protonated (Figure 5B1), causing the polymer to carry positive charges. With the increase of pH, the polymer became less positively charged due to deprotonation until pH 5.8 where PAN became neutrally charged (Figure 5B-2). Further increase in pH caused the polymer to be negatively charged, possibly due to competitive specific adsorption of OH- anions on both imine and amine functional groups (Figure 5B-3). A similar interpretation was proposed by Zhang and Bai (41) in their study of the altered electrokinetic properties of polypyrrole, where negative ζ potentials at elevated solution pHs were attributed to the specific binding of OH- to the nitrogen functional groups on the polymer chain.

FIGURE 5. (A) Electrophoretic mobility of polyaniline before (PAN) and after (PAN-Hg) Hg(II) sorption in 0.01 M NaNO3 solution. PAN-Hg was prepared by equilibrating PAN with Hg(II) solution (pH ) 5.5, ionic strength ) 0.2 M NaNO3, C0 of Hg(II) ) 100 mg/L). (B) The proposed physicochemical properties of polyaniline in aqueous solutions as a function of solution pH. 5226

9

ENVIRONMENTAL SCIENCE & TECHNOLOGY / VOL. 43, NO. 14, 2009

Discussion The observed impact of solution pH on Hg(II) uptake could be explained by changes in both the physicochemical properties of PAN (Figure 5B) and Hg(II) speciation (calculated from MINEQL+ as demonstrated in the Supporting Information, Figure S2A). At low pHs less than 4.0, nitrogen atoms of the imine functional groups were competitively bound by protons, causing decrease in mercury adsorption. It is likely that the decreased adsorption for mercury was primarily ascribed to the competitive binding between protons and mercury for the imine functional groups, although columbic repulsion might also be responsible for the decreased Hg(II) adsorption at these low pHs (15). In addition, in very acidic solution, Hg2+(aq) becomes the dominant mercury species (Supporting Information, Figure S2A). Compared with Hg(OH)+ and Hg(OH)2 species of relatively smaller hydrated sizes, Hg2+(aq) has lower affinity to nitrogen-containing binding sites (15, 21), causing further decrease in mercury adsorption. With increasing pH, protons are released from the imine groups (Figure 5B-1-2), leaving more binding sites available for mercury adsorption. Also, as pH increases, Hg(OH)+ and Hg(OH)2 become the dominant species (e.g., Hg(OH)2 at pH 5.5), resulting in increased mercury removal. The observed decrease in mercury removal above pH 6.0 could be primarily attributed to the changed physiochemical properties (Figure 5B-3). Chloride and humic acid were observed to inhibit Hg(II) sorption on PAN because they served as competitive ligands to complex mercury. As for Cl-, it could change the distribution of Hg(II) species by competing with OH- to form other mercury complexes (Figure S2, Supporting Information) that have lower affinity to the binding sites than Hg-OH complexes, causing lower Hg(II) removal. Similar phenomena have been reported for the impact of Cl- on Hg(II) sorption by bacteria (42) and soil components (18). As for humic acid, its inhibitory effect on mercury adsorption was suggested to be attributed to its nitrogen and sulfur-containing ligands that have strong affinity to complex mercury. Although the concentration of sulfurcontaining ligands might be low for humic acid (e.g., 0.15-0.4 mol of total S/kg soil humic acid (24)), they have been ascribed as the main binding sites in view of their extremely high affinity toward Hg(II) (24, 25). Two Hg4f7/2 doublet-peaks appeared on the highresolution XPS of mercury sorbed polyaniline (Figure 4B). The absence of Hg4f7/2 peaks at BEs lower than 100.6 eV suggested that no elemental mercury (Hg(0)) was formed by PAN during adsorption as BE for Hg4f7/2 of Hg(0) is around 99.9 eV (43). However, as BEs for Hg(I) and Hg(II) are very close (23), we are not able to differentiate these two mercury species in our Hg4f XPS spectra if they do

exist simultaneously. At this point, we believe that Hg(II) removal by polyaniline is caused primarily by formation of complexes between Hg(II) and the nitrogen-containing functional groups on the polymer matrix. In previous studies, Hg(II) was also deemed as the primary form for mercury adsorption on polyaniline (33). It needs to be mentioned that although the Hg4f7/2 peak at 101.1 eV was within the domain of typical BE of 99.9-101.6 eV (43), the peak at 103.0 eV was definitely outside this range. This could be caused by the existence of charged imine functional groups on PAN polymer matrix due to protonation. In their study of mercury sorption on pyrite, Ehrhardt et al. (22) also attributed the two Hg4f7/2 peaks to mercury sorption at two different zones, where the peak at 100.7 eV was assigned to pyrite zones, while that at 103.8 eV to Fe(III) oxyhydroxides patches having different charge properties from the bulk FeS2. The existence and characteristics of two Hg4f doublets (Hg1 and Hg2) in the high-resolution XPS (Figure 4B) suggested that at pH 5.5 mercury could be complexed to two types of general binding sites (Figure 5B-1 and 2) on the polymer chain. At this pH, the imine nitrogen was preferably protonated over the amine nitrogen, which is harder to be protonated than imine. This is further supported by the nitrogen speciation from N1s XPS spectra of PAN-Hg and PAN 5.5 (Figure 4C-b and c), where imine, amine and protonated imine coexisted at pH 5.5. It is possible that nitrogen atoms from imine, which had relatively higher affinity to mercury, served as type-one binding sites (e.g., for Hg1), while nitrogen atoms from both amine and protonated imine served as type-two mercury binding sites (Hg2) because of their similar lower strength to complex mercury. The two-site assumption above can qualitatively explain our desorption data. As shown in Table 1, 1.0 M HCl desorbed 23.5% mercury from Hg1 sites, but 58.3% from sites corresponding to Hg2, suggesting qualitatively that Hg(II) has higher affinity to the type-one binding sites. On the basis of this study, a schematic mechanism for mercury uptake by PAN was proposed. Adsorption of Hg(II) by polyaniline from water solution was due to complexation between mercury and the nitrogen-containing binding sites on the polymer matrix. Many factors including physicochemical properties of PAN and water chemistry conditions (e.g., solution pH and competitive ligands) have impacts on adsorption. The impact of pH on mercury was possibly due to the formation of protonated imine functional groups on the polymer matrix at low pH, which had relatively weak affinity to mercury. At higher pH, mercury adsorption could be inhibited by competitive adsorption of OH- on both imine and amine functional groups. Solution pH may also impact adsorption through Hg(II) speciation. Specifically, at pH 5.5, multiple adsorption sites consisting of imine, amine and protonated imine coexisted. These sites could be further subcategorized into type-one (imine) and type-two (amine and protonated imine) sites, with type-one sites having higher affinity to mercury than type-two sites. Both chloride and humic acid can inhibit mercury adsorption by competitive complexation, with the effect of Cl- being much more pronounced than the latter. Future studies by advanced instrumentation methods (e.g., XANES), however, are needed to verify the proposed mechanism for Hg(II) adsorption on polyaniline.

Acknowledgments Funding for this research was provided by the National Natural Science Foundation of China (Grants 20540420542, 20377021, 40473053) and by the Ministry of Education of China under Grant 106081. We thank graduate students Guanghui Chen and Jun Yin for their experimental assistance

and Rong Tang for helping us to synthesize polyaniline in the initial stage of this project. The anonymous reviewers are gratefully acknowledged for their valuable comments and suggestions that help to substantially improve this manuscript.

Supporting Information Available Figures for FTIR spectroscopy of the as-synthesized PAN and Hg(II) speciation simulated by MINEQL+. This information is available free of charge via the Internet at http:// pubs.acs.org.

Literature Cited (1) Pacyna, E. G.; Pacyna, J. M. Global emission of mercury from anthropogenic sources in 1995. Water Air Soil Pollut. 2002, 137, 149–165. (2) Nam, K. H.; Gomez-Salazar, S.; Tavlarides, L. L. Mercury(II) adsorption from wastewaters using a thiol functional adsorbent. Ind. Eng. Chem. Res. 2003, 42, 1955–1964. (3) Wang, Q.; Kim, D.; Dionysiou, D. D.; Sorial, G. A.; Timberlake, D. Sources and remediation for mercury contamination in aquatic systemssA literature review. Environ. Pollut. 2004, 131, 323–336. (4) Aguado, J.; Arsuaga, J. M.; Arencibia, A. Adsorption of aqueous mercury(II) on propylthiol-functionalized mesoporous silica obtained by cocondensation. Ind. Eng. Chem. Res. 2005, 44, 3665–3671. (5) Zhang, L.; Planas, D. Biotic and abiotic mercury methylation and demethylation in sediments. Bull. Environ. Contam. Toxicol. 1994, 52, 691–698. (6) Hintelmann, H.; Keppel-Jones, K.; Douglas Evans, R. Constants of mercury methylation and demethylation rates in sediments and comparison of tracer and ambient mercury availability. Environ. Toxicol. Chem. 2000, 19 (9), 2204–2211. (7) Cameron, R. E. Guide to Site and Soil Description for Hazardous Waste Site Characterization. Volume 1: Metals. U.S. Environmental Protection Agency: Washington, DC, 1992. (8) U.S. Environmental Protection Agency. National Primary Drinking Water Standards; EPA 816-F-01-007; EPA: Washington, DC, 2001. (9) Namasivayam, C.; Kadirvelu, K. Uptake of mercury(II) from wastewater by activated carbon from an unwanted agricultural solid by-product: coirpith. Carbon 1999, 37, 79–84. (10) Budinova, T.; Savova, D.; Petrov, N.; Razvigorova, M.; Minkova, V.; Ciliz, N.; Apak, E.; Ekinci, E. Mercury adsorption by different modifications of furfural adsorbent. Ind. Eng. Chem. Res. 2003, 42, 2223–2229. (11) Feng, X.; Fryxell, G. E.; Wang, L. Q.; Kim, A. Y.; Liu, J.; Kemner, K. M. Functionalized monolayers on ordered mesoporous supports. Science 1997, 276, 923–926. (12) Mercier, L.; Pinnavaia, T. J. A functionalized porous clay heterostructure for heavy metal ion (Hg2+) trapping. Microporous Mesoporous Mater. 1998, 20, 101–106. (13) Guerra, D. L.; Airoldi, C.; Viana, R. R. Performance of modified montmorillonite clay in mercury adsorption process and thermodynamic studies. Inorg. Chem. Commun. 2008, 11, 20–23. (14) Zhu, X.; Alexandratos, S. D. Polystyrene-supported amines: Affinity for mercury(II) as a function of the pendant groups and the Hg(II) counterion. Ind. Eng. Chem. Res. 2005, 44, 8605–8610. (15) Das, S. K.; Das, A. R.; Guha, A. K. A Study on the adsorption mechanism of mercury on Aspergillus versicolor biomass. Environ. Sci. Technol. 2007, 41, 8281–8287. (16) Li, N.; Bai, R.; Liu, C. Enhanced and selective adsorption of mercury ions on chitosan beads grafted with polyacrylamide via surface-initiated atom transfer radical polymerization. Langmuir 2005, 21, 11780–11787. (17) Vieira, R. S.; Beppu, M. M. Interaction of natural and crosslinked chitosan membranes with Hg(II) ions. Colloids Surf., A. 2006, 279, 196–207. (18) Yin, Y. J.; Allen, H. E.; Li, Y.; Huang, C. P.; Sanders, P. F. Adsorption of mercury(II) by soil: Effects of pH, chloride, and organic matter. J. Environ. Qual. 1996, 25, 837–844. (19) Tiffreau, C.; Lu ¨tzenkirchen, J.; Behra, P. Modeling the adsorption of mercury(II) on (hydr)oxides. I. Amorphous iron oxide and quartz. J. Colloid Interface Sci. 1995, 172, 82–93. (20) Bonnissel-Gissinger, P.; Alnot, M.; Lickes, J. P.; Ehrardt, J. J.; Behra, P. Modeling the adsorption of mercury(II) on (hydr)oxides. II. FeOOH (geothite) and amorphous silica. J. Colloid Interface Sci. 1999, 215, 313–322. (21) Wu, P.-Y. Aqueous mercury partition and modeling. PhD. Thesis, University of Utah, Salt Lake City, UT, 2003. VOL. 43, NO. 14, 2009 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

9

5227

(22) Ehrhardt, J. J.; Behra, P.; Bonnissel-Gissinger, P.; Alnot, M. XPS study of the sorption of Hg(II) onto pyrite FeS2. Surf. Interface Anal. 2000, 30, 269–272. (23) Behra, P.; Bonnissel-Gissinger, P.; Alnot, M.; Revel, R.; Ehrhardt, J. J. XPS and XAS study of the sorption of Hg(II) onto pyrite. Langmuir 2001, 17, 3970–3979. (24) Hesterberg, D.; Chou, J. W.; Hutchison, K. J.; Sayers, D. E. Bonding of Hg(II) to reduced organic sulfur in humic acid as affected by S/Hg ratio. Environ. Sci. Technol. 2001, 35, 2741–2745. (25) Ravichandran, M. Interactions between mercury and dissolved organic mattersA review. Chemosphere 2004, 55, 319–331. (26) Shimano, J. Y.; MacDiarmid, A. G. Polyaniline, a dynamic block copolymer: Key to attaining its intrinsic conductivity. Synth. Metal. 2001, 123, 251–262. (27) Kumar, S.; Verma, R.; Gangadharan, S. Application of poly(aniline) as an ion-exchanger for the separation of palladium, iridium, platinum, and gold prior to their determination by neutron-activation analysis. Analyst 1993, 118, 1085–1087. (28) Sahayam, A. C. Determination of Cd, Cu, Pb, and Sb in environmental samples by ICP-AES using polyaniline for separation. Fresenius’ J. Anal. Chem. 1998, 362, 285–288. (29) Sofiane, B.; Didier, H.; Laurent, L. P. Synthesis and characterization of composite Hg-polyaniline powder material. Electrochim. Acta 2006, 52, 62–67. (30) Gupta, R. K.; Shankar, S. Toxic waste removal from aqueous solutions by polyaniline: A radiotracer study. Adsorp. Sci. Technol. 2004, 22, 485–496. (31) Gupta, R. K.; Singh, R. A.; Dubey, S. S. Removal of mercury ions from aqueous solutions by composite of polyaniline with polystyrene. Sep. Purif. Technol. 2004, 38, 225–232. (32) Balarama Krishna, M. V.; Karunasagar, D.; Rao, S. V.; Arunachalam, J. Preconcentration and speciation of inorganic and methyl mercury in waters using polyaniline and gold trap-CVAAS. Talanta 2005, 68, 329–335. (33) Remya Devi, P. S.; Kumar, S.; Verma, R.; Sudersanan1, M. Sorption of mercury on chemically synthesized polyaniline. J. Radioanal. Nucl. Chem. 2006, 269, 217–222.

5228

9

ENVIRONMENTAL SCIENCE & TECHNOLOGY / VOL. 43, NO. 14, 2009

(34) Lu, Q. F.; Huang, M. R.; Li, X. G. Synthesis and heavy-metal-ion sorption of pure sulfophenylenediamine copolymer nanoparticles with intrinsic conductivity and stability. Chem.sEur. J. 2007, 13, 6009–6018. (35) Wang, P. C.; Venancio, E. C.; Sarno, D. M.; MacDiarmid, A. G. Simplifying the reaction system for the preparation of polyaniline nanofibers: Re-examination of template-free oxidative chemical polymerization of aniline in conventional low-pH acidic aqueous media. Reac. Funct. Polym. 2009, 69 (4), 217–223. (36) Yan, X.; Han, Z.; Yang, Y.; Tay, B. Fabrication of carbon nanotubepolyaniline composites via electrostatic adsorption in aqueous colloids. J. Phys. Chem. C. 2007, 111, 4125–4131. (37) Zhang, G.; Sun, Sh.; Yang, D.; Dodelet, H. P.; Sachera, E. The surface analytical characterization of carbon fibers functionalized by H2SO4/HNO3 treatment. Carbon 2008, 46, 196–205. (38) Walcarius, A.; Delacote, C. Mercury(II) binding to thiolfunctionalized mesoporous silicas. Anal. Chim. Acta 2005, 547, 3–13. (39) Kang, E. T.; Neoh, K. G.; Tan, K. L. Polyaniline: A polymer with many interesting intrinsic redox states. Prog. Polym. Sci. 1998, 23, 277–324. (40) Menardo, C.; Nechtschein, M.; Rousseau, A.; Travers, J. P. Investigation on the structure of polyaniline: 13C NMR and titration studies. Synth. Met. 1988, 25, 311–322. (41) Zhang, X.; Bai, R. Surface electric properties of polypyrrole in aqueous solutions. Langmuir 2003, 19, 10703–10709. (42) Daughney, C. J.; Siciliano, S. D.; Rencz, A. N.; Lean, D.; Fortin, D. Hg(II) adsorption by bacteria: A surface complexation model and its application to shallow acidic lakes and wetlands in Kejimkujik National Park, Nova Scotia, Canada. Environ. Sci. Technol. 2002, 36, 1546–1553. (43) Hutson, N. D.; Attwood, B. C.; Scheckel, K. G. XAS and XPS characterization of mercury binding on brominated activated carbon. Environ. Sci. Technol. 2007, 41, 1747–1752.

ES803710K