Distribution and Early Diagenesis of Antimony Species in

the Periodic Table and can exist in four oxidation states (-III,. 0, III, V) although only the two ... in the 1970s to approximately 6.3 nowadays. McF...
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Environ. Sci. Technol. 2003, 37, 1163-1168

Distribution and Early Diagenesis of Antimony Species in Sediments and Porewaters of Freshwater Lakes YU-WEI CHEN,† TIAN-LONG DENG,‡ MONTSERRAT FILELLA,§ AND N E L S O N B E L Z I L E * ,† Department of Chemistry & Biochemistry, Laurentian University, Sudbury, Ontario, P3E 2C6 Canada, Department of Applied Chemistry, Chengdu University of Technology, Chengdu, 610059 China, and Department of Inorganic, Analytical and Applied Chemistry, University of Geneva, 30 Quai Ernest Ansermet, CH-1211 Geneva 4, Switzerland

The study identifies the role played by different components of natural aquatic systems on the poorly known geochemistry of antimony. Different chemical forms of antimony were measured in porewaters and sediments of two Sudbury lakes characterized by contrasting redox conditions at the sediment-water interface. In porewaters, Sb(III) was present under reducing conditions where it could exist as SbS2- according to thermodynamic calculations. Sb(V) was detected mainly under oxic and mildly reducing environments where its presence was attributed to the oxidizing effect of iron and manganese oxyhydroxides or to the slow kinetics of reduction by dissolved sulfide or possible complexation by it. A third form of Sb identified as refractory was obtained after UV irradiation of the water samples, suggesting an association of Sb to low molecular weight natural organic matter. The distribution of Sb in sediments of the two lakes revealed (through the comparison of profiles and statistical correlations) the importance of iron and manganese oxyhydroxides in controlling the behavior of Sb, particularly in the lake where the interface was clearly oxic. Porewater profiles indicate that the dissolution of manganese and iron oxyhydroxides under anoxic conditions leads to the simultaneous release of dissolved Sb previously sorbed onto those compounds. In reducing sediments, the control of the solubility of Sb by iron sulfides is suggested.

Introduction Antimony is broadly distributed in the environment as a consequence of natural processes and human activities which include the combustion of fossil fuels and mining and smelting operations. Antimony belongs to the group 15 of the Periodic Table and can exist in four oxidation states (-III, 0, III, V) although only the two higher states are most frequently encountered in environmental samples. Antimony and its compounds are considered as pollutants of priority interest by the United States Environmental Protection Agency (1) and the European Union (2). The toxic action of Sb species is less than that of As except for organic Sb * Corresponding author phone: (705)675-1151 ext. 2114; fax: (705)675-4844; e-mail: [email protected]. † Laurentian University. ‡ Chengdu University of Technology. § University of Geneva. 10.1021/es025931k CCC: $25.00 Published on Web 02/15/2003

 2003 American Chemical Society

compounds which are more toxic than the analogous As compounds. Sb(III) is more toxic than the pentavalent form (3). In an extensive review (4), concentrations of Sb in freshwaters were reported ranging from a few ng/L (pM) to a few µg/L (nM) depending on location. The mean concentration of Sb in oceans is around 200 ng/L or 1.6 nM where Sb is mainly present in the pentavalent state (3). Sb(III) has been detected in surface waters where its presence seems to be mainly due to phytoplanktonic activity (5-8). Sb(III) is the main Sb species in anoxic waters (9-11) and porewaters (9). Except for highly contaminated areas, Sb concentrations in sediments are in the order of a few µg/g or less. Although Sb has occasionally been reported to be bound to reducible iron and aluminum oxides in sediments (12, 13) or to organic carbon (14), the geochemical behavior of Sb in sediments is still largely unknown. A slight enrichment of Sb in crusts of manganese and iron oxides has been reported in sediments of Lake Baikal (15). Amorphous forms of iron and manganese oxyhydroxides were shown to adsorb Sb(III) strongly and oxidize it effectively into Sb(V) in a few days (16). Redox equilibria are not always observed between Sb(III) and Sb(V) species. Apart from the previously mentioned presence of Sb(III) in surficial oxic waters, that of Sb(V) in anoxic environments has also been reported and attributed to the formation of thioantimony complexes (6, 17, 18). A better understanding of the geochemical behavior of Sb in the aquatic environment requires information on the speciation of this element in porewaters and its distribution in the sediment. In this paper, the detailed distribution of Sb species in porewaters and sediments of two Sudbury area freshwater lakes was obtained by using diffusion porewater samplers and sequential extraction techniques. Sb species were analyzed by Hydride Generation Atomic Fluorescence Spectrometry. The sediment-water interface (SWI) of the two lakes differed in pH and redox conditions, and the objective of the study was to look at the specific influence of the two parameters on the geochemical cycling of Sb in freshwater sediments.

Experimental Section Sampling Sites. Sediment and porewater samples were collected from two freshwater lakes located in Sudbury in August 2000. Clearwater Lake (46°22′ N; 81°03′ W), identified as CLW in this paper, was strongly acidified in the past by atmospheric emissions, and its pH has changed from to 4.2 in the 1970s to approximately 6.3 nowadays. McFarlane Lake (46°25′ N, 80° 57′ W), or MCF, a well buffered slightly alkaline lake at pH ) 7.5, is located only 5 km away from CLW and receives similar atmospheric loading of trace elements from the Sudbury smelters (see ref 19 for more details). For both lakes, samples were collected at littoral sites of approximately 7 m in depth. These two contaminated lakes were selected for this study based on criteria of acidification level and oxic status at the sediment-water interface; CLW being well oxygenated all year and the isolated basin in MCF becoming close to anoxia in summer months. A previous study on Se suggested that the SWI of MCF was characterized by more reducing conditions (lower dissolved oxygen levels) as compared to that of CLW (20). Both lakes become stratified during the summer months. Water and Sediment Sampling. Profiles of temperature, pH, and dissolved oxygen (DO) in both lakes were obtained with a YSI 3800 Water Quality Monitor in August 2000. Porewater samples were collected in one occasion using in situ diffusion samplers (porewater equilibrators) that had been intensively deoxygenated for 72 h before insertion in VOL. 37, NO. 6, 2003 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

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the sediment by a diver. They were allowed to equilibrate with interstitial waters for 14 days (20, 21). Each sampler contains two rows of 3.3-mL chambers disposed to obtain a 1-cm resolution in depth. Two series of two collectors were attached back to back and inserted side by side in order to collect larger volumes of interstitial water at each corresponding depth. Sediments at both locations are very soft. After equilibration, samplers were retrieved from the sediment by the diver, and water samples of the same level were collected and pooled after piercing the 0.2 µm filtration membrane with the tip of a micropipet. Separate measurements on dissolved Fe and Mn in porewater collected from opposite and lateral chambers did not show differences in concentrations larger than 10%. A first series of samples was immediately collected for dissolved sulfide analysis and transferred to a flask containing the fixing agent for sulfide. Other samples were transferred into precleaned and preacidified Teflon bottles (except for pH measurements). Samples were kept on ice during the transfer to the laboratory where they were immediately frozen at -80 °C to prevent any bacterial or chemical alteration before analysis. A subsample was used to measure the pH close to the sediment-water interface. Two undisturbed sediment cores were collected by the diver with a lightweight corer at each site close to the porewater sampling location. After retrieval, cores were immediately transported to the laboratory for extrusion under N2 atmosphere. Cores were sliced into 0.5cm sections for the top 5 cm and 1.0-cm sections for the rest of the core. The subsamples were placed in polyethylene bottles and frozen at -80 °C until analysis. Sample Preparation and Analysis. To limit the contact with air and loss of volatile sulfide, the first porewater subsamples were collected for dissolved sulfide. A 2.0-mL volume was pipetted in a tube already containing the aminesulfuric acid reagent to fix the dissolved sulfide (22). The tube was capped and brought to the laboratory where the rest of the procedure was completed. The absorbance of the complex was read on a UV-visible spectrophotometer at 670 nm the same day. It had been previously verified that the blue-color complex was stable for at least 3 days. A 1.0-mL subsample was used to measure pH close to the SWI using a needlelike combined pH electrode. Antimony species were analyzed by continuous flow Hydride Generation Atomic Fluorescence Spectrometry (HG-AFS) on a PSA-10-055 Millenium Excalibur system according to a procedure previously developed (23). Hydride compounds were atomized in a miniature H2 flame produced by the reaction between HCl and alkaline sodium tetrahydroborate solution. For the determination of Sb(III), a subsample of preacidified porewater was analyzed the next day in 3 M HCl Trace Metal Grade. To measure Sb(V), another aliquot was treated with a 1.2 M KI solution to reduce all inorganic Sb to Sb(III) before determination by HG-AFS in 5 M HCl. Sb(V) was then obtained by difference. A third aliquot was treated with 2% (v/v) HNO3 and submitted to UV photooxidation (300 nm) in a Rayonet photochemical chamber for 5 h to destroy the dissolved organic matter and determine the fraction of Sb associated to it. A diluted solution of 8-hydroxyquinoline was added to control chemical interferences caused by heavy metals and mask the undesired fluorescence emission signal due to Sb(V) during the determination of all Sb species (23). Duplicate or triplicate measurements were done regularly in all series. After the determination of Sb by HG-AFS, the concentration of organic/refractory Sb was obtained by subtracting the sum of Sb(III) and Sb(V) from the total Sb concentration (Sb-Tot). The detection limit of the instrument for Sb analysis was at 5 ng/L or 0.1 nM, and the precision was usually 5% on triplicate samples. Total dissolved Fe and Mn were determined by flame and graphite furnace AAS. 1164

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Dissolved organic carbon (DOC) was determined using a DOC analyzer based on persulfate UV oxidation. The water content of sediments was determined on ovendried samples at 105 °C for 4 h until constant weight. To estimate the fraction of total organic carbon (TOC) in sediments, a known mass (∼0.5 g) of oven-dried (40 °C) sediment was subjected to a temperature of 750 °C in a muffle furnace for 4 h. The TOC content was obtained from the difference of sample weight before and after ashing and expresses the mass. The procedure used (24) to analyze the total reducible sulfur (TRS), defined as the sum of elemental sulfur, pyritic sulfur, and acid volatile sulfide (AVS), is designed to convert TRS into H2S in a hot acidic Cr2+ solution in the presence of ethanol. The H2S gas is collected as a ZnS precipitate and ultimately analyzed by iodimetry. A sequential extraction method was used to determine the proportion of total Sb associated to oxidized and organic fractions of the sediment. This is a modified version of the Tessier et al. method (25) and has also been used to study the geochemistry of Se (20, 26). The selectivity of sequential extraction has been extensively discussed in the literature, but the approach remains useful especially when dealing with elements present at trace concentrations. Wet sediment samples were submitted to sequential extraction by (a) a 0.2 M oxalic acid solution buffered to pH 2 with ammonium oxalate for 8 h to remove Sb mainly bound to iron and manganese oxyhydroxides and identified as Sb-Oxal and (b) an acidic solution (pH ) 2) of H2O2 (9% v/v) at 85 °C for 5 h to remove Sb bound to organic matter and amorphous sulfides in the sediment and identified as Sb-H2O2. The selectivity of these extraction methods will be discussed later. The Sb species in both extracts were converted to Sb(III) using KI (final concentration 2% w/v) and 8-hydroxyquinoline (0.1%) to eliminate interference before analysis by HG-AFS (23). The corresponding fractions of oxyhydroxides (Fe-Oxal; Mn-Oxal) and bound to organic matter and amorphous sulfides (Fe-H2O2; Mn-H2O2) Fe and Mn in the sediments were obtained from the same extractions and determined by flame AAS. For the determination of total concentrations of Sb, Fe, and Mn, all sediment samples were digested in a mixture of acids in Teflon bombs according to the procedure described in refs 20 and 23. The quality of the digestion and analysis was controlled through repeated determinations of two certified standard sediment (PACS-1 and BCSS-1) from the National Research Council of Canada. Statistical correlations (r) were calculated between concentrations of dissolved species in porewater and between fractions in sediments.

Results and Discussion pH and Redox Status of the Sediment-Water Interface. The two littoral stations of CLW and MCF showed different profiles of DO. The water column of CLW was well oxygenated, whereas DO decreased strongly from 8.5 to less than 6.0 mg/L close to the SWI of MCF. The Winkler method was used to measure DO in water samples collected by a diver just above the sediment; DO was around 9.0 mg/L in CLW and less than 5.0 mg/L in MCF. However the real DO concentration of the water located just above the SWI cannot be obtained by those techniques if only a very thin layer of overlying water is depleted in oxygen as it seems to be the case in MCF (see discussion below). Both the YSI probe and the diver can contaminate a thin anoxic layer by introducing oxygenated overlying water. The two lakes also differed in their pH values; in CLW, pH ranged from 6.2 near the surface to 5.5 close to the SWI. In MCF, it varied from 8.3 to 7.2. There was a good agreement between pH values measured at the bottom of the lakes and those measured in peepers for overlying waters. Values were almost constant at pH 7.0 in MCF but increased from 5.0 in the lake water to 6.6 in the

FIGURE 1. Concentrations of dissolved (a) Mn, (b) Fe, and (c) sulfide across the sediment-water interface of Clearwater and McFarlane Lakes.

FIGURE 2. Concentrations of dissolved Sb species across the sediment-water interface of (a) Clearwater and (b) McFarlane Lakes. sediment porewater of CLW. Temperature trends were similar in the two lakes. The pH and redox status at the SWI of the two lakes and the corresponding effects on the geochemical behavior of Se were reported in a previous study (20). Several measured parameters corroborated the differences in pH and redox conditions existing at the SWI of the two lakes. The SWI was clearly oxic and acidic in CLW, whereas the SWI of MCF showed depleted values of DO and neutral pH conditions. Consequences of those conditions on some dissolved constituents are presented in Figures 1 and 2. Relatively reducing conditions at the SWI of MCF are confirmed by significant peaks of dissolved Mn, dissolved sulfide (Figure 1a,c), and Sb(III) (Figure 2b) in overlying waters. Even at the neutral pH conditions of MCF, concentrations of dissolved Fe (Figure 1b) were also relatively high likely due to the low dissolved oxygen levels at the SWI that do not favor the kinetics of precipitation of iron oxyhydroxides (20). Relatively high concentrations of dissolved Mn and Fe close to the oxic SWI of CLW (Figure 1a,b) might be related to the more acidic conditions of this lake. However, in contrast to MCF, concentrations of Sb(III) in overlying waters of CLW were close to or below the detection limit of the analytical technique (Figure 2a). In the absence of kinetic considerations, the presence of dissolved Sb(III) only depends on the redox status of the waters (at the range of concentrations considered), so its presence at the SWI of MCF can be taken

as a direct effect of the more reducing conditions of SWI in this lake in comparison to CLW. Sediments can also be used to corroborate the interplay of pH and redox conditions at the SWI. The surficial enrichment of iron and manganese oxyhydroxides in CLW (Figure 3a,b) is characteristic of an oxygenated SWI and is in contrast with the situation in the neutral but more reducing SWI of MCF where the surficial enrichment is not obvious (Figure 3d,e). Contrary to CLW, the presence of an oxidized layer is not visually detectable in MCF. It is particularly interesting to analyze the evolution of Mn oxyhydroxides at the same location of CLW as the pH continued to rise from 4.3 in the 1980s to approximately 5.5 in 2000. The concentration of Mn-Oxal in surficial sediments was 1.0 µmol/g in the early 1980s (27) and 4.0 µmol/g in 1995 (20) and reached 25 µmol/g in 2000. Assuming relatively constant levels of dissolved oxygen through these years, the Mn-Oxal increase would be mainly due to the increasing pH conditions of the lake, now more favorable to the kinetics of oxidation of Mn2+ in the lake waters which can in turn limit its diffusion from the sediment to the water column (28). The Effect of Iron and Manganese Oxyhydroxides on Sb. The distribution of total Sb in CLW showed a subsurface maximum peak just below the SWI followed by a sharp decrease to the background level of 1-2 nmol/g. The portion of Sb associated to oxides, Sb-Oxal, represented around half of the total in the surficial enrichment (Figure 4a). The profile of total Sb in MCF also showed a maximum peak between 4 and 5 cm in depth (Figure 4b), and a large portion of the total Sb was associated to the oxide phase. The affinity of Sb for iron and manganese oxides has been shown in some adsorption studies for Sb(III) (16, 29, 30) and Sb(V) (31-33). However, when translating this type of laboratory results to real systems, it is important to keep in mind that competition for adsorption of elements on natural oxyhydroxides is high, particularly in contaminated environments. When elements are present at very low levels, as it is the case for Sb in the lakes studied, competition with other metal and metalloid ions for adsorption on iron and manganese oxides (in nonacidic lakes) cannot be neglected, but it is difficult to estimate. In the case of Sb, a few sequential extraction studies on sediments also show a favorable association of Sb with Fe and/or Al compounds (12, 13, 29). Statistical correlations (r) between Sb-Tot or Sb-Oxal and Fe-Tot or Fe-Oxal are stronger than those between Sb and Mn fractions (Supporting Information). This seems to be in contradiction with a recent study on sediments of Lake Baikal where total Sb was often found to correlate better with Mn VOL. 37, NO. 6, 2003 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

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FIGURE 3. Distribution of (a, d) Mn, (b, e) Fe, and TOC (c, f) fractions in sediments of Clearwater (top) and McFarlane (bottom) Lakes.

FIGURE 4. Distribution of Sb forms in sediments of (a) Clearwater and (b) McFarlane Lakes. in distinct Mn crusts than with Fe in Fe crusts (15). In Lake Baikal however the two layers identified as Mn and Fe crusts were clearly separated in the sediments and this allowed a separate comparison of the two layers with the Sb present in the solid phase. Because the two oxides coexist in the same sediment layer in the Sudbury lakes, it becomes much more difficult to determine whether there is a preferential association of Sb with any of the two oxides, especially when considering that concentrations of iron oxyhydroxides are 1166

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40-50 times higher in surficial sediments than their Mn counterpart. In MCF significant correlations were also obtained between Sb-Tot or Sb-Oxal and Fe-Tot or FeOxal. The preference of Sb for either manganese or iron oxyhydroxides does not appear clearly when profiles are compared (Figures 3 and 4). It might be explained by the poor oxidizing conditions existing at the SWI of MCF that do not favor the existence of manganese or iron oxyhydroxides (Figure 3d,e), The unusual presence of Mn-Ox or Fe-Ox in deeper reducing sediments might be explained by the existence of carbonate forms of Mn such as rhodochrosite or Fe as siderite, particularly in the circumneutral MCF lake (19). Those two mineral forms could be easily dissolved during the extraction by the oxalic buffer. Direct comparison between the profiles of dissolved Fe, Mn, and inorganic Sb in porewaters and statistical correlations (Supporting Information) also corroborates the affinity of Sb for iron and manganese oxyhydroxides, particularly in CLW. The simultaneous increase of dissolved Fe, Mn, and Sb in the porewaters of this lake (maximum values around 3-4 cm) can be related to the dissolution of iron and manganese oxyhydroxides under reducing conditions, with the concomitant release of the adsorbed Sb (Figures 1 and 2). Profiles of total and inorganic Sb both suggest an upward diffusion of dissolved Sb and readsorption/reprecipitation close to the SWI. The similarity between dissolved Sb and dissolved Fe was less obvious in MCF porewaters (Figures 1 and 2 and Supporting Information). In this lake, dissolved Mn present in larger concentrations than dissolved Fe seemed to follow more closely Sb profiles (better correlations) than

dissolved Fe did. Although, the results obtained point to a significant control of Sb fate by iron oxyhydroxides in the sediments of the study lakes, the affinity of Sb for iron oxides seems not to be as strong as that of As (34). The Effect of Natural Organic Matter. Very few studies provide experimental evidence for complexation of Sb with natural organic matter (NOM) in natural waters (35). In a study involving voltametric measurements, a significant fraction of Sb released after the UV irradiation of seawater was attributed to NOM complexes (36). By using XAD resins, a stable association between aquatic humics and Sb was identified in the bottom waters of Lake Pavin, France (37). More recently, a fraction of Sb associated with NOM (ranging from 35 to 67% of total Sb) was identified in three Sudbury lakes (23). The same analytical procedure was used in the present work, and it was found that 35-85% (average 61%) of Sb in porewaters of CLW and 35-55% (average 50% Sb) in those of MCF were present in a refractory fraction which could be partially or entirely made of NOM. In CLW, good statistical correlations were observed between DOC and SbTot (r ) 0.86), Sb(III) (r ) 0.74), and Sb-inorg (r ) 0.78). Correlations between DOC and Sb species were poorer in MCF. The fact that the presence of refractory or NOM-bound Sb is not often reported in the literature could be due to the analytical protocols applied. Most protocols are usually designed to measure Sb(III) and total Sb separately, then the difference between the two is considered as Sb(V), but they do not include any treatment for destruction of NOM. In our study, the collection of porewaters was done using porewater collectors with water samples diffusing through a 0.22 µm membrane. It is therefore possible that dissolved or colloidal Sb complexed to low molecular weight NOM crosses the membrane and is released into the solution as free Sb after the UV treatment. It is known that the exposure to UV radiation can destroy some dissolved organic matter present in natural waters thus liberating trace metals bound to it (38-40). In the sediments of both lakes, the fraction extracted by an acidic solution of H2O2 and identified as Sb-H2O2 was always very small (Figure 4) and not correlated to TOC. The significant proportion of Fe extracted by the same H2O2 solution from the sediments of McFarlane could indicate the presence of a distinct mineral phase such as amorphous FeS and will be discussed below. The Effect of Sulfide. The binding of Sb to sulfide ions has been studied extensively in the laboratory, and a large variety of sulfur-containing compounds of Sb, mainly stable under strong alkaline conditions, has been reported in the literature (see ref 35 for a detailed discussion). However, it is difficult to predict what happens in reducing sediments where competitors for sulfide ions such as Fe2+ exist in much larger concentrations than Sb and where heterogeneous patchy pH and redox microenvironments are known to exist. Strong statistical correlations exist between Sb-Tot or SbOxal and total reducible sulfur, TRS (r ) 0.88 and 0.87, respectively), measured in sediments of MCF where dissolved sulfide and TRS were much more abundant than in CLW. The positive correlations could be indicative of a control of solubility of Sb by sulfide, but thermodynamic calculations using the Jess program (41) predict the formation of the soluble form SbS2- under reducing conditions. Therefore, the solubility of Sb is more likely controlled by sorption on iron sulfide since Fe-Tot is also highly correlated with TRS (r ) 0.86). Calculations (42) also predict that amorphous iron sulfide can be formed in sediments of MCF but not likely in CLW (Supporting Information). It is known that the Fe-H2O2 fraction (Figure 3b,e) can also contain some amorphous sulfide (25, 27), and this can explain the high correlation existing between Fe-H2O2 and TRS (r ) 0.93). It is also noticeable in Figure 4b that a significant portion of

Sb was associated to a nondefined fraction of the total. The fraction obtained after the subtraction of Sb-Oxal and SbH2O2 from Sb-Tot could correspond to Sb associated to sulfides as this fraction is indeed highly correlated with TRS (r ) 0.80). Amorphous iron sulfide and pyrite are known to play an important role in controlling the solubility of other metalloids such as As (43) and Se (20). The presence of Sb(V) in surficial sediments of both lakes where sulfides were measurable (Figures 1c and 2) might be related to slow kinetics of reduction or to the coexistence of iron and manganese oxyhydroxides with dissolved sulfide. These oxyhydroxide compounds are known to efficiently oxidize Sb(III) into Sb(V) in natural sediments (16). The presence of Sb(V) and dissolved sulfide above the SWI of MCF where dissolved oxygen was depleted but still present suggests the existence of the form SbS43- as predicted by thermodynamic calculations (41) (Supporting Information). In the presence of dissolved sulfide, the formation of SbS43is particularly favored at pH values between 6 and 8. Other authors have reported the existence of Sb(V) in sulfidic waters (9, 17) or explained the presence of measurable dissolved sulfide in oxygenated waters by the formation of relatively stable metal sulfide complexes (44, 45). Recently, the formation of Sb(V) thioanions upon dissolution of stibnite (Sb2S3) in deoxygenated aqueous NaHS solutions was confirmed by X-ray absorption spectroscopy (46). The authors invoked the promoted oxidation of Sb(III) to Sb(V) by HS- to explain this result. When Sb profiles of the two lakes are compared, it is obvious that sediments of MCF had a more efficient retention capacity for Sb over the last century than those of CLW. In the latter, Sb is concentrated near the SWI and involved in a dynamic cycle with the precipitation-dissolution-diffusionreprecipitation of iron and (more recently) manganese oxyhydroxides. This cycle is favored by the oxic conditions prevailing in the lake and close to the SWI although the historical acidic conditions of the past decades did not promote the stability of oxyhydroxide (particularly Mn) compounds. Under oxidizing and acidic conditions, thermodynamic calculations predict the dominance of the negatively charged Sb(OH)6- (Supporting Information), which can then adsorb more efficiently on positively charged amorphous iron oxyhydroxides. On the other hand, mildly reducing conditions at the SWI of MCF did not facilitate the establishment of the same dynamic cycle with oxyhydroxides and rather promoted the retention of sulfur through the formation of iron sulfide and pyrite, which could in turn retain Sb by sorption or favor Sb complexation. Profiles of dissolved species and distributions in the solid phase show that the early diagenesis of Sb in freshwater sediments is affected by several factors and compounds. Apart from the expected influence of thermodynamics (oxidation state) and kinetics, the mobility of Sb in sediments likely depends on complexation with natural organic matter and with sulfides. The sorption of Sb on iron and manganese oxyhydroxides under oxidizing conditions and that on iron sulfides in reducing sediments, which also implies a role of the lake/porewater pH, must be also considered.

Acknowledgments This work received financial support from the Natural Sciences and Engineering Research Council of Canada, the Elliot Lake Research Field Station of Laurentian University, and Questron Technologies Corporation. Technical assistance from Sanford Clark is acknowledged.

Supporting Information Available Figures A-D and Tables 1-4. This material is available free of charge via the Internet at http://pubs.acs.org. VOL. 37, NO. 6, 2003 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

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Received for review July 2, 2002. Revised manuscript received January 7, 2003. Accepted January 20, 2003. ES025931K