Environ. Sci. Technol. 2004, 38, 4535-4541
Antimony(III) Binding to Humic Substances: Influence of pH and Type of Humic Acid JOHANNA BUSCHMANN* AND LAURA SIGG EAWAG (Swiss Federal Institute for Environmental Science and Technology), P.O. Box 611, CH-8600 Du ¨ bendorf, Switzerland
Conditional distribution coefficients (Dom) for Sb(III) binding to three commercial humic acids (terrestrial, coal, and aquatic) were measured at environmentally relevant Sb(III)/DOC ratios and as a function of pH using an equilibrium dialysis method. Maximum binding of Sb(III) was observed around pH 6 for two of the humic acids. The third humic acid showed constant Dom values up to pH 6 and decreasing Dom values for pH > 6. Sb(III)/DOC ratio was found to be important for Dom (20 times higher Dom for 60 times lower Sb(III)/DOC ratio). Moreover, Dom depends on the individual humic acid, suggesting that different functional groups are involved and/or different degrees of stabilization by chelation or H-bridges. Chemical modeling of Sb(III)humics binding at different pH values is consistent with two binding sites involving (i) a phenolic entity forming a neutral complex and (ii) a carboxylic entity forming a negatively charged complex. Under environmentally relevant conditions, over 30% of total Sb(III) may be bound to natural organic matter.
Introduction Antimony is used in flame retardants, batteries, and alloys and is exposed to the environment as Sb(0) or Sb(III). In environmental samples, it exists mainly as Sb(III) and Sb(V). The solution chemistry of antimony has been reviewed by Fillela et al. (1). According to thermodynamic data, antimony should exist as Sb(V) in oxic systems and as Sb(III) in anoxic ones (2, 3). The standard reduction potential (E°) of Sb(OH)6-/ Sb(OH)4- is 0.238 V (4), and for a typical ratio of Sb(V)/Sb(III) ([Sb(OH)6-] ) 10-8 M, [Sb(OH)3] ) 10-10 M, pH 7) p is 0.425. The toxicity of antimony is assumed to be similar to that of arsenic with respect to effects and mechanism (5, 6); therefore, studying mobility and bioavailability of antimony in aquatic systems is important for risk assessment. The environmental fate of antimony could be affected by natural organic matter (NOM), which has an impact on its speciation. In the pH range of 2-11, Sb(III) should form a neutral complex, Sb(OH)3, whereas Sb(V) should exist as a negatively charged complex, Sb(OH)6- (pKa([Sb(H2O)(OH)2)]+) ) 1.2; pKa(Sb(OH)3) ) 11.8) (7). It is probable that the neutral species Sb(OH)3 interacts more strongly with the negatively charged NOM than Sb(OH)6-. Indeed, considering stability constants of small organic ligands suggests that Sb(III)-NOM complexes may be of importance (7, 8). Despite the known relevance of NOM on Sb speciation in aquatic systems, * Corresponding author phone: +41 1 823 50 86; fax: + 41 1 823 53 11; e-mail:
[email protected]. 10.1021/es049901o CCC: $27.50 Published on Web 07/20/2004
2004 American Chemical Society
experimental information on Sb binding to NOM is scarce (9, 10). Binding constants under environmentally relevant conditions are not available, so the distribution of Sb in the environment based on calculations cannot be predicted. In this study, we determined conditional distribution coefficients (Dom) for Sb(OH)3 and three commercial humic acids that differed in carbon content and number of functional groups: a terrestrial, a coal, and an aquatic humic acid. An equilibrium dialysis method was used. The pH dependence as well as the influence of [Sb(III)]o/DOC ratio were fitted with a discrete site model. Using this model, Sb(III) speciation in aquatic systems may be predicted in the presence of NOM.
Materials and Methods Reagents. Standard solutions of Sb(III) were prepared by dilution of Sb2O3, 1002 ( 2 mg/L, in 2 M HCl from Merck (Switzerland). The mineral acid HCl (30%) was of suprapure grade (Merck). Sodium azide (purum p.a.) and sodium chloride solution (5 M; 6 (Figure 4). For AHA, at a [Sb(III)]o/DOC ratio of 1.36 nmol/mg of DOC, lower Dom values are observed than at 0.838 nmol/mg (Figure 2). Stronger binding sites seem to be involved at lower [Sb(III)]o/DOC ratios. This tendency is confirmed by plotting Dom as a function of the nonbound Sb(III), Cw (Figure 5), where nonbound means not bound to AHA and may include complexes with chloride or buffer which are, however, negligibly weak (14). At pH 6.1, as shown in this example, the
lower the [Sb(III)]o/DOC ratio, the stronger the binding of Sb(III) to AHA. Dom increases by a factor of about 20 when the [Sb(III)]o/DOC ratio is decreased by a factor of 60. Quantifying the occupation of binding sites for AHA results in 435 µmol of Sb(III)/mol total functional groups for a [Sb(III)]o/DOC ratio of 7.57 nmol/mg of DOC (pH 6.1, saturation conditions), meaning that 0.5‰ of all binding sites are occupied (Figure 5). For AHA, a mole fraction of 36 µmol of Sb(III)/mol of total functional groups is found for a [Sb(III)]o/DOC ratio of 2.74 nmol/mg of DOC (pH 6.1), while under the same conditions, SRHA has a mole fraction of 250 µmol of Sb(III)/mol of total functional groups (Figure 3). Thus, for the same [Sb(III)]o/DOC ratio and pH, seven times higher mole fractions are found for SRHA than AHA. SRHA exhibits more functional groups per gram of DOC (factor ≈ 2) and has smaller molecules (factor ≈ 2, as large molecules hinder the diffusion of Sb(III) to the binding sites), which would lead to about 4 times higher Dom values. Our findings of 10-15 times higher Dom values and 7 times higher mole fractions of SRHA as compared to AHA suggest differences in the sort of the functional groups involved in Sb(III) binding. A 7 times higher affinity of SRHA for Sb(III) might be due to thiol functional groups that are known to bind Sb(III) quite strongly (15-17), due to binding pockets facilitating chelation or H-bridges and/or due to metals involved in binding Sb(III). Comparison of AHA and LHA at a [Sb(III)]o/DOC ratio of 1.36 and 1.61 nmol/mg of DOC (pH 6.1) gives mole fractions of 68 and 37 µmol of Sb(III)/mol of total functional groups, respectively. LHA exhibits larger molecules than AHA (factor ≈ 2) and has a similar number of functional groups per VOL. 38, NO. 17, 2004 / ENVIRONMENTAL SCIENCE & TECHNOLOGY
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kilogram of DOC. These findings suggest that AHA exhibits binding sites that are stronger or better available, as diffusion of Sb(OH)3 into buried sites of the large LHA molecules might be hindered. Binding Mechanisms. The influence of pH on the Sb(III)-humic acid distribution can be explained by the acid/ base characteristics of the humic acids (Table 1) and a discrete site complexation mechanism. At pH 6.1, a substantial part of the carboxylic functional groups is deprotonated. Two complexation mechanisms are proposed: (i) ligand exchange at the Sb center and release of one (eq 2) or two (eq 3) hydroxides, respectively (R represents the large rest of the humics) and (ii) formation of a negatively charged complex (eq 4). In addition, chelation, H-bridges or cationic metals may stabilize Sb(III) bound to humics.
For AHA and SRHA, which exhibit maximum binding at pH 6, competition by H+ for binding sites is significant at lower pH values, while increasing concentrations of OH- at pH > 6 shift the equilibria (eqs 2 and 3) to the nonbound Sb(III). Equilibrium 4, however, is only pH dependent with respect to the deprotonation of the binding sites involved but is not influenced by increasing OH- for pH > pKa + 1 of these binding sites. Thus, the fraction of the negatively charged complex increases up to pH ) pKa + 1 of the corresponding functional group and then levels off. For LHA, complex formation with release of at least one OH- is probably more important than formation of a negatively charged complex, as our findings of almost constant Dom values at pH < 6 suggest (see also Model for Interaction of Sb(III) with Humic Acid section). Comparison of the Π-donor characteristics of R-COO-, phenolic Ar-O-, and HO- shows that aryl-oxy should stabilize the positively charged Sb center the most effectively (RCOO- < HO- < Ar-O-) (18) and therefore facilitate release of OH- (eqs 2 and 3), while R-COO- would rather form a negatively charged complex (eq 4). Consequently, carboxylic functional groups are proposed to be involved in binding mechanism ii (eq 4) whereas phenolic entities are assumed to bind Sb(III) with binding mechanism i (eqs 2 and 3). In addition, interactions between Sb(OH)3 and cationic metals, such as Fe(III) or Al(III), could be involved in the association, analogously to As-NOM complexation (19). Although Sb exhibits a stronger cationic character than As, binding of either the free electron pair at the Sb center or an oxygen bridge to the cationic metal is plausible. As metal 4538
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concentrations are rather low as compared to the total number of functional groups in the humics studied here (1‰-1%), however, we assume that the contribution of this binding mechanism to the total Sb(III) complexation is rather small. Moreover, complexation of Sb(III) with a cationic metal presupposes decomplexation of the involved metal, at least partially, from the humic material. Sb(OH)3 with its cationic character is probably not able to compete with carboxylic or phenolic functional groups in binding cationic metals. In addition, if binding to cationic metals such as Fe(III) were to be a dominant mechanism underlying the association, Dom values would be expected to increase for pH < 6 because of the Fe(III) speciation as a function of pH. This does not correspond with our findings of decreasing Dom values (AHA, SRHA) and constant Dom values (LHA) with decreasing pH for pH < 6. In any case, the quantification of metal-Sb(III) binding in the presence of humic acid exhibiting much more binding sites is difficult. Complexation versus Hydrophobic Binding. Although Sb(OH)3 is a neutral complex, hydrophobic interactions with humic material should only play a minor role. Dom values of neutral pollutants with SRHA are reported to be 10-20 times smaller than with AHA (20, 21). Because of its small size, SRHA is to a lesser degree able to adopt molecular conformation for hydrophobic interactions with neutral pollutants than AHA. Host-guest interactions involving n-alkanes, fatty acids, or benzenecarboxylic acids and humic material were explained by the small molecules becoming entangled with the larger molecules so that separation by dialysis was not possible (22). As a matter of fact, observation of 10-15 times higher Dom values for SRHA than for AHA at the same [Sb(III)]o/DOC ratio is a clear hint for a specific complexation of Sb(III) by the corresponding functional groups. Model for Interactions of Sb(III) with Humic Acid. Conditional distribution coefficients were measured for two sets of conditions: (i) continuously varying pH at constant [Sb(III)]o/DOC ratio (Figures 2-4) and (ii) continuously varying [Sb(III)]o/DOC ratios at constant pH (Figure 5). With a discrete site model written in Microsoft Excel, Sb(III) binding to humics was simulated by a two binding site model with two different functional groups and two different complex formation reactions. The pKa values of the sites (pKi and pKy), the number of functional groups per gram of DOC, and the association constant for Na+ to these sites were given by Arnold et al. (23) and Westall et al. (24). pKi and pKy correspond to pKa3 and pKa1, respectively (Table 1). KNa is assumed to be identical for all sites (24). The model involved the following equilibria with their corresponding equilibrium constants:
LiH ) Li- + H+ Ki
(5)
LyH ) Ly- + H+
Ky
(6)
Li- + Na+ ) LiNa
KNa
(7)
Ly- + Na+ ) LyNa
KNa
(8)
Li- + Sb(OH)3 ) LiSb(OH)2 + OH- Ki′ Ly- + Sb(OH)3 ) LySb(OH)3-
Ky′
(9) (10)
The equilibrium concentrations of LiSb(OH)2 and LySb(OH)3were calculated according to
[LiSb(OH)2] )
Ki′[Li-]([Sb]tot - [LySb(OH)3-]) [OH-] + Ki′[Li-]
(11)
and
[LySb(OH)3-] )
Ky′[Ly-]([Sb]tot - [LiSb(OH)2]) 1 + Ky′[Ly-]
(12)
where [Li-] and [Ly-] are the fractions of deprotonated ligands calculated according to
[Li-] )
1 ‚[LiH]tot 1 + 10pKi-pH + KNa[Na +]
(13)
and the assumption that [Li-] . [LiSb(OH)2] and [Ly-] . [LySb(OH)3-]. DOC-normalized Dom values were calculated by
Dom )
[LiSb(OH)2] + [LySb(OH)3-] [Sb(OH)3][HA]{C}
[L kgDOC-1]
(14)
where [Sb(OH)3] ) [Sb]tot - [LiSb(OH)2] - [LySb(OH)3-], [HA] ) concentration of humics (in kg/L), and {C} ) carbon content of the corresponding humic acid (in kg/kg). Minimization of ∑i(Domi,simulated - Domi,measured)2 for continuously varying pH at constant [Sb(III)]o/DOC ratio was done by adjusting the complex formation constants Ki′ and Ky′ using a curve fit procedure of Excel. As can be seen in Figures 2-4 (dotted lines), agreement between model and data was good. Whereas for AHA and SRHA both binding mechanisms are important, for LHA mechanism i (eqs 2 and 3) predominates. The values Ki′ and Ky′ determined from the pH-dependent data were then applied to the fixed-pH, varying [Sb(III)]o/ DOC ratios data without further adjustment. Here, correspondence was fairly good only because the model underestimates binding for low Sb(III)/DOC ratios and overestimates it for high ratios. Obviously, low concentrations of high energy sites and high concentrations of low energy sites are involved in Sb(III) binding. A Freundlich isotherm plot on the basis of the data in Figure 5 has a slope of 0.41, indicating that Sb(III) is bound to weaker sites at higher Sb(III)/DOC ratios. However, including different concentrations of different energy sites in the model requires more data. This model is thus valid under limited conditions, only for certain Sb(III)/DOC ratios. Still, the highly complicated interactions of Sb(III) with humic acids can be described quite well with this simple model involving only two sites and two binding mechanisms. Oxidation of Sb(III) Bound to Humics. We found evidence that humic acid catalyzes Sb(III) oxidation. When the inside and outside solution of the dialysis tube were separated after 25 d of equilibration (pH 8.4), analysis of Sb(III) and Sb(V) showed an Sb(III)/Sb(V) ratio of 0.24 for the humic-acid containing solution, whereas the outside solution (without humics) had an Sb(III)/Sb(V) ratio of 0.62. Recovery of total Sb was 75 ( 2%. Moreover, in the distribution experiments the sorption capacity reached after 7 d of equilibration remains constant for the next 18 d and for all pH values tested (Figure 6a). The nonbound Sb(III) concentration (Cw) however, slowly decreases during the same time (Figure 6b). This loss depends on the [Sb(III)]o/DOC ratio. For five different [Sb(III)]o/DOC ratios (0.131-7.57 nmol/mg of DOC), Cs is constant during the period studied (25 d), whereas Cw is decreasing (Figure 7 in Supporting Information). The smaller the [Sb(III)]o/DOC ratio, the faster the decrease of Cw. However, without humic acid (blank), no significant decrease of Cw is observed for 65 d. The observed constant Cs implies that Sb, once oxidized, is released again into the solution so that another Sb(III) can bind to the sites involved in the oxidation step. Such an oxidation, for example, by disulfides or quinones that are inherent structures of humic material, is accompanied by
FIGURE 6. Cs (a) and Cw (b) as a function of time for four different pH values: pH 4.6 (b), pH 6.1(O), pH 7.2 (1), and pH 8.4(0). OH- binding of the resulting octahedral Sb(V) center (eq 15). After oxidation, Sb(V) is assumed to be released as Sb(OH)6-.
As Sb(III) is slowly oxidized in the presence of humic acids, Dom values observed after 7 d of equilibration are actually pseudo-equilibrium coefficients. Nevertheless, the Dom values give quantitative information about the affinity of Sb(III) for humic material. Comparison with Literature. Bowen et al. reported binding of Sb(III) to a commercial peat humic acid (9). By working with a radio tracer method and elution from a Sephadex G-75 column, they found two Sb(III)-humic complexes. A total of 75% of Sb(III) was bound to a size fraction of about 8000 (a) and 25% to a size fraction of about 1000 (b). At pH 9-11, fraction b increased, whereas at pH 2-4 both complexes were destroyed. A quantitative study on Sb-humic interactions was published by Pilarski et al. (10). Using a terrestrial humic acid, they found saturation capacities of 23 µmol/g when a solution of Sb(OH)3 was equilibrated with the humic material. For a solution of potassium antimonyl tartrate, a saturation capacity of 53 µmol/g was found. In the work presented here, a saturation capacity of 2.3 µmol/g of AHA was found. It is noteworthy, however, that all experiments were performed in the presence of 0.05 M VOL. 38, NO. 17, 2004 / ENVIRONMENTAL SCIENCE & TECHNOLOGY
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NaCl. Na+ competes with Sb(III) for binding sites (see Table 1). For SRHA a saturation capacity of 33 µmol/g of SRHA is obtained, which is in the same order of magnitude as the findings by Pilarski et al. (10). Metal-to-dissolved organic matter (DOM) concentration ratios can be very important in studying metal-binding to humic substances. At low Hg(II)/DOM ratios, Haitzer et al. (25) observed strong interactions of Hg(II) and DOM, indicative of mercury-thiol bonds. At high Hg(II)/DOM ratios, however, significantly smaller Dom values indicate mercury-oxygen bonds (25). In the case of Sb(III) binding to the humic acids studied here, Dom values also depend on the Sb(III)/DOC ratio. At very low [Sb(III)]o/DOC ratios, Sb(III)-thiol bonds might be involved, which are known to be stronger than Sb(III)-oxygen bonds (26, 27, 15, 17). For the effective removal of Sb(III) from industrial wastewater, a thiolcontaining chelating resin has been developed (28). Moreover, not only monodentate but also multidentate complexes might be formed. Humics offer binding pockets by H-bridges analogously to proteins. Binding of Sb(OH)3 that exhibits three OH- ligands may easily be involved in H-bridges. Finally, cationic metals might be involved in Sb(III) binding as discussed above. With respect to pH dependence, Sb(III)-binding behavior to humics was similar to that of tributyltin with AHA (23). Dom exhibited a maximum close to the acidity constant (pKa) of tributyltin. As tributyltin is positively charged at pH < 6.2, inner sphere complex formation was postulated between the tin atom and the deprotonated functional groups. For pH > 6.2, the tributyltin is uncharged (binding of hydroxide). In contrast to Sb(III) binding to humics, however, hydrophobic interactions were postulated to explain the decrease in Dom with increasing pH. Binding of cations such as Ca2+, Cu2+, and Cd2+ typically increases with increasing pH (29). Co2+ binding to LHA, for example, exhibited an increase in Dom by a factor of 20 when pH was increased from 4.8 to 6.9 (24). As OH- acts as a ligand in Sb(OH)3, however, OHcompetes with the deprotonated binding sites in Sb(III) binding, which explains the observed weaker sorption at higher pH values. Sorption studies of Sb(III) to mineral surfaces reported decreasing amounts of Sb(III) sorbed to hydrous oxides of Mn, Fe, and Al at pH > 6 (30). Moreover, at higher pH values, oxidation of Sb(III) to Sb(V) is facilitated when Sb(III) is bound to mineral surfaces (31). Whereas oxidation of Sb(OH)3 by O2 is only significantly fast at pH > 11 (4), the oxidation of Sb(III) bound to mineral surfaces takes place at pH < 11. The higher electron density at the bound Sb(III) center provided by electron-donating groups facilitates the oxidation (31). As NOM is reported to act as an oxidizing or reducing agent in the redox chemistry of arsenic, an oxidation of Sb(III) in the presence of humic material acting as a redox agent or catalyst is possible as well (19). Stability constants of Sb(III) with small organic ligands support the proposed binding mechanisms for the interaction of Sb(III) with humic entities. Carboxylic functional groups as found in acetic acid or EDTA form stable Sb(III) complexes (7). Hydrolysis of diacetato-Sb(III) complex is important at pH > 8 (1 mM acetate) (8). The complex SbEDTA- is hydrolyzed at pH > 6 (1 mM HEDTA3-) (7). Vicinal phenolic functional groups in ligand L as found in catechol or tiron form stable bidentate complexes Sb(OH)L and are not hydrolyzed even at pH > 10 (7). The mixed bidentate complex Sb(III)-2,3-dihydroxybenzoic acid is stable in the whole pH range (26). Moreover, cysteine-containing peptides bind Sb(III) effectively. Although the exact speciation of Sb(III) is not explicitly mentioned, studies for trypanothione and glutathione report that Sb(III)-thiol bonds are involved in complex formation (15-17). These publications clearly show that Sb(III) is bound to carboxylic, phenolic and thiol groups 4540
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at environmentally relevant pH conditions. Direct comparison of the equilibrium constants Ki′ and Ky′ determined here with those of known Sb(III) complexes is not possible because of stoichiometric differences in the corresponding association reactions. For an equivalent number of functional groups per liter and same pH, however, the calculated fraction bound to humics can be compared with that bound to small organic ligands. Thus, for about 70 µM functional groups and pH 7, which corresponds to 5 mg/L DOC of SRHA, 30% of 8.21 nM Sb(OH)3 is bound to humics while 99% is bound to catechol (vicinal phenolic functional groups) and 6.5% is bound to acetate (8). The hypothesis of a complexation mechanism of Sb(III) to the functional groups of humic acids is thus supported by the stability of Sb(III) with small organic ligands. Environmental Considerations. When studying speciation of antimony in the aquatic environment, binding of Sb(III) to NOM has to be considered. Although antimony is found abundantly in its pentavalent form, Sb(OH)6-, Sb(III) is thermodynamically stable under anoxic conditions (2) and is therefore important in risk assessment and mobility considerations of antimony. At environmentally relevant conditions ([DOC] ) 5 mg/L; pH 7; [Sb(OH)3] ) 8.21 nM (1 µg/L), I ) 0.05), up to 0.74 µmol of Sb(III)/g of DOC (90 µg of Sb(III)/g of DOC) are bound to aquatic humic acids, meaning that 30% of total Sb(III) are bound to humics. At smaller ionic strengths, it is expected that even more Sb(III) is bound to humics. Such large fractions of bound Sb(III) should not be neglected in mobility considerations nor with respect to the reported oxidation of Sb(III) to Sb(V) catalyzed by humic material.
Acknowledgments We thank Adrian Ammann for help with the experimental setup, Se´bastien Meylan for help with voltammetry, AnnKathrin Leuz for analysis with hydride generation atomic fluorescence spectroscopy, and David Kistler for help in laboratory work. Adrian Ammann, Silvio Canonica, Annette Johnson, Ann-Kathrin Leuz, and three anonymous reviewers are kindly acknowledged for reviewing the manuscript.
Supporting Information Available One figure showing Cs and Cw/Cwo as a function of time for different [Sb(III)]o/DOC ratios. This material is available free of charge via the Internet at http://pubs.acs.org.
Literature Cited (1) Filella, M.; Belzile, N.; Chen, Y.-W. Earth-Sci. Rev. 2002, 59, 265-285. (2) Andreae, M. O.; Froelich, P. N. Tellus 1984, 36, 101-117. (3) Filella, M.; Belzile, N.; Chen, Y.-W. Earth-Sci. Rev. 2002, 57, 125-176. (4) Leuz, A.-K. Diploma Work, Du ¨ bendorf, 2002. (5) Cai, J.; Salmon, K.; DuBow, M. S. Microbiology 1998, 144, 27052713. (6) Gebel, T. Chem.-Biol. Interact. 1997, 107, 131-144. (7) Smith, R. M.; Martell, A. E. Critical Stability Constants; Plenum Press: New York, 1976; Vol. 4: Inorganic Complexes. (8) Wojtas, R. Rocz. Chem. 1975, 49, 1231-1237. (9) Bowen, H. J. M.; Page, E.; Valente, I.; Wade, R. J. J. Radioanal. Nucl. Chem. 1979, 48, 9-16. (10) Pilarski, J.; Waller, P.; Pickering, W. Water Air Soil Pollut. 1995, 84, 51-59. (11) Quentel, F.; Filella, M. Anal. Chim. Acta 2002, 452, 237-244. (12) Andreae, M. O.; Asmode´, J.-F.; Foster, P.; Van’t dack, L. Anal. Chem. 1981, 53, 1766-1771. (13) Deng, T.-L.; Chen, Y.-W.; Belzile, N. Anal. Chim. Acta 2001, 432, 293-402. (14) Filella, M.; May, P. M. Geochim. Cosmochim. Acta 2003, 67, 4013-4031. (15) Sun, H.; Yan, S.; Cheng, W. S. Eur. J. Biochem. 2000, 267, 54505457. (16) Sun, H.; Yan, S.; Ding, K. J. Inorg. Biochem. 2001, 86, 446.
(17) Yan, S.; Ding, K.; Zhang, L.; Sun, H. Angew. Chem., Int. Ed. 2000, 39, 4260-4262. (18) Angst, W.; Diem, D.; Meister, E.; Uhlig, W.; Togni, A. Chemie I+II; Course script, ETH Zurich, 2003. (19) Redman, A. D.; Macalady, D. L.; Ahmann, D. Environ. Sci. Technol. 2002, 36, 2889-2896. (20) Chin, Y. P.; Aiken, G. R.; Danielsen, K. M. Environ. Sci. Technol. 1997, 31, 1630-1635. (21) Chiou, C. T.; Kile, D. E.; Brinton, T. I.; Malcolm, R. L.; Leenheer, J. A.; Maccarthy, P. Environ. Sci. Technol. 1987, 21, 1231-1234. (22) Smeulders, D. E.; Wilson, M. A.; Kannangara, K. G. S. Org. Geochem. 2001, 32, 1357-1371. (23) Arnold, C. G.; Ciani, A.; Muller, S. R.; Amirbahman, A.; Schwarzenbach, R. P. Environ. Sci. Technol. 1998, 32, 29762983. (24) Westall, J. C.; Jones, J. D.; Turner, G. D.; Zachara, J. M. Environ. Sci. Technol. 1995, 29, 951-959. (25) Haitzer, M.; Aiken, G. R.; Ryan, J. N. Environ. Sci. Technol. 2003, 37, 2436-2441. (26) Anderegg, G.; Malik, S. Helv. Chim. Acta 1970, 53, 577-600. (27) Anderegg, G.; Malik, S. Helv. Chim. Acta 1970, 53, 564-569.
(28) Uchiumi, A.; Tokunaga, S.; Eyama, T. Removal of heavy metals from industrial wastewaters development of new chelating resins for arsenic(III), antimony(III) and copper(II) ions; First International Conference on Processing Materials for Properties, Osaka, Japan, 1993. (29) Benedetti, M. F.; Milne, C. J.; Kinniburgh, D. G.; Van Riemsdijk, W. H.; Koopal, L. K. Environ. Sci. Technol. 1995, 29, 446-457. (30) Thanabalasingam, P.; Pickering, W. F. Water Air Soil Pollut. 1990, 49, 175-185. (31) Belzile, N.; Chen, Y.-W.; Wang, Z. Chem. Geol. 2001, 174, 379387. (32) Chin, Y. P.; Aiken, G.; O’Loughlin, E. Environ. Sci. Technol. 1994, 28, 1853-1858. (33) Chin, Y. P.; Gschwend, P. M. Geochim. Cosmochim. Acta 1991, 55, 1309-1317. (34) Beckett, R.; Jue, Z.; Giddings, J. C. Environ. Sci. Technol. 1987, 21, 289-295.
Received for review January 20, 2004. Revised manuscript received May 13, 2004. Accepted June 8, 2004. ES049901O
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