Depositional Influences on Porewater Arsenic in Sediments of a

Aug 21, 2008 - a Mining-Contaminated Freshwater. Lake. GORDON TOEVS, †. MATTHEW J. ... porewater As concentrations within Lake CDA sediments...
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Environ. Sci. Technol. 2008, 42, 6823–6829

Depositional Influences on Porewater Arsenic in Sediments of a Mining-Contaminated Freshwater Lake GORDON TOEVS,† M A T T H E W J . M O R R A , * ,† LEIGH WINOWIECKI,† DANIEL STRAWN,† MATTHEW L. POLIZZOTTO,‡ AND SCOTT FENDORF‡ Division of Soil & Land Resources, PO Box 442339, University of Idaho, Moscow, Idaho 83844-2339, and Department of Geological and Environmental Sciences, Stanford University, Stanford, California 94305-2115

Received April 03, 2008. Revised manuscript received June 13, 2008. Accepted June 23, 2008.

Arsenic-containing minerals mobilized during mining activities and deposited to Lake Coeur d’Alene (CDA), Idaho sediments represent a potential source of soluble As to the overlying water. Our objective was to delineate the processes controlling porewater As concentrations within Lake CDA sediments. Sediment and porewater As concentrations were determined, and solid-phase As associations were probed using X-ray absorption near-edge structure (XANES) spectroscopy. Although maximum As in the sediment porewaters varied from 8.4 to 16.2 µM, As sorption on iron oxyhydroxides at the oxic sediment-water interface prevented flux to overlying water. Floods deposit sediment containing variable amounts of arsenopyrite (FeAsS), with major floods depositing large amounts of sediment that bury and preserve reduced minerals. Periods of lower deposition increase sediment residence times in the oxic zone, promoting oxidation of reduced minerals, SO42- efflux, and formation of oxide precipitates. Depositional events bury oxides containing sorbed As, transitioning them intoanoxicenvironmentswheretheyundergodissolution,releasing As to the porewater. High Fe:S ratios limit the formation of arsenic sulfides in the anoxic zone. As a result of As sequestration at the sediment-water interface and its release upon burial, decreased concentrations of porewater As will not occur unless As-bearing erosional inputs are eliminated.

Introduction Lake Coeur d’Alene (CDA) is a natural lake of glacial origin located in northern Idaho. Lake inputs include the St. Joe River that lies within a relatively pristine watershed and the CDA River, the receiving water body for the South Fork of the CDA River drainage within which is located a world-class mining district (1). As a result of this activity, mine tailings and mill slurries contaminated with Pb, As, Cd, Zn, and other metal(loid)s have accumulated throughout the floodplain of the CDA River and in the sediments of Lake CDA. Annual * Corresponding author phone: 208-885-6315; fax: 208-885-6315; e-mail [email protected]. † University of Idaho. ‡ Stanford University. 10.1021/es800937t CCC: $40.75

Published on Web 08/21/2008

 2008 American Chemical Society

and episodic flood events continue to resuspend, transport, and redeposit these sediments both in the river and in the lake (2). Estimates indicate that contamination within the lake sediments exceeds 75 million metric tons and covers 85% of the lakebed (3); however, except for Zn, the overlying water column typically meets U.S. Environmental Protection Agency standards for primary contact, recreation (1). This is not the case in the sediment porewaters where dissolved trace metals such as Cu, Pb, and Zn are found at concentrations that exceed the criterion maximum concentration (CMC) or the criterion continuous concentration (CCC) (1, 4). Approximately 11 500 t of As has been transported to Lake CDA (3), most of it originating from sulfidic minerals exposed during mining activities (5). Soluble As released by oxidation of reduced As minerals typically exists as the As(III) oxyanion arsenite and the As(V) oxyanion arsenate. Arsenite is considered more toxic (6), having adverse effects on aquatic species at concentrations of 0.25 µM (7). Both arsenate and arsenite sorb to iron and manganese oxyhydroxides (8–10), limiting their diffusion into overlying water (11, 12). Arsenite will also precipitate as an arsenic sulfide in anaerobic zones if adequate sulfide is available (13, 14). Redox transformations and adsorption processes have been identified as potential contributors to As geochemical cycling in CDA sediments (15). Increased concentrations of soluble As at redox boundaries have been observed in stratified water columns and sediments (16–18). Analysis of Lake CDA sediment porewaters indicates that microbially mediated redox processes change from oxic/suboxic to anoxic at sediment depths of approximately 10 cm as indicated by decreases in sulfate concentrations (19). Significant nitrate is only found in porewaters at the oxic sediment-water interface and soluble Mn and Fe below sediment depths of approximately 3 cm. Geochemical changes in As species are expected to occur in this dynamic region since the elements of interest in As cycling (i.e., Fe and S) are redox sensitive. Also relevant are the frequent flood events that bury oxic materials, transitioning them to a depth where reductive dissolution occurs (19). Prior research on As in Lake CDA sediments has been limited to quantifying total As in the sediments (3, 20) and laboratory studies to determine potential As release under changing redox conditions (15, 21). Unfortunately, we lack a comprehensive understanding of the processes governing As cycling within these sediments, thus limiting our predictive capabilities and ability to make informed management decisions. Our objective was to delineate the depositional and redox processes controlling porewater As within Lake CDA sediments by quantifying solid-phase and aqueous concentrations and chemically characterizing sediment As distributions.

Materials and Methods Sample Sites. Two sample sites, Harlow Point (HP) and Peaceful Point (PP), were established in the southern portion of Lake CDA within 1 km of the mouth of the CDA River in an area contaminated by heavy metal(loid)s (3, 20). Our sites were chosen so that previously collected stratigraphic data (3) could be utilized in data interpretation. Core 123 taken from this part of Lake CDA by Horowitz et al. (3) corresponds to the area of our HP site and core 6 to our PP site. The water depths at PP and HP averaged 19 and 15 m, respectively. A map can be found elsewhere showing that HP is located closer to the mouth of the CDA River than PP (19). VOL. 42, NO. 18, 2008 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

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Materials. All chemicals were reagent grade and used without further purification. All sample containers and labware were cleaned with 2% HNO3 and rinsed with deionized water prior to use. Analytical standards and quality assurance standards for ion chromatography (IC) and inductively coupled plasma (ICP) spectroscopy were purchased from Spex CertiPrep (Metuchen, NJ) (19). Sampling Devices. Porewater was collected in situ utilizing Plexiglas equilibrium dialysis samplers as previously described (19). The assembled dialyzer was submerged in a container filled with deionized H2O continuously purged with N2 for 48 h prior to deployment to prevent O2 introduction to the sediments (22). The dialyzers were transported in N2purged H2O and only removed at the time they were handed to the diver. Cores were collected in polycarbonate tubes 5 cm in diameter and 50 cm in length (19). Sample Retrieval and Preservation. Divers inserted the dialyzers vertically into the sediment in April 2002 where they were allowed to equilibrate for 4 weeks. Upon retrieval, porewater for anion analysis was removed from the dialyzer cells and preserved in a sample vial under anaerobic conditions at 4 °C. Porewater for cation analysis was placed in a sample vial and preserved with HNO3. Cores obtained at the same time were sealed to prevent leaks and oxidation, packed in an airtight container with ice packs, and purged with N2 (19). In the laboratory, sediment cores were transferred to a N2-filled glovebox and extruded from the coring tubes with care taken to avoid compaction and elongation. Cores were subsectioned and homogenized samples placed in vials that were frozen until analysis by X-ray absorption near-edge structure (XANES) spectroscopy. Analytical Methods. Sediment samples for total metal(loid)s were analyzed as previously described (19). Samples were microwave-digested with a 9:2:3 ratio of HNO3, HCl, and HF in accordance with EPA method 3052 (23) and quantified using a Thermal Jarrell Ash ICP-AES. Total carbon in sediment samples was determined by dry combustion at 1100 °C on an Elementar VarioMax elemental analyzer (Elementar, Hanau, Germany). Organic carbon in the sediment was measured after inorganic carbon was removed by treatment with 6 M HCl containing 3% FeCl2 · 4H2O (24). The reported values are the average of duplicate samples. Sulfate was measured by ion chromatography, and soluble Fe and Mn were quantified with ICP-AES using described methods (19). Porewater As was analyzed on a HP 4500 ICPMS with a flow rate of 0.5 mL min-1 after a 5:1 dilution accomplished by a peristaltic pump. In order to avoid airborne contamination, sample preparation was performed in a positive pressure clean hood and the autosampler was housed in an enclosure protected from contamination by a high efficiency particulate air (HEPA) filter. Detection limits for As using ICP-MS were calculated as three times the standard deviation of the blank. Reported concentrations of porewater As are the average of three replicate measurements of each sample. Particle size analysis was performed on subsamples of core sections dried at 90 °C until a stable weight was achieved. The dry samples were cooled in a vacuum desiccator and stored until particle size was analyzed using a procedure that included sample dispersion with sodium hexametaphosphate and sonification (25). Fraction totals exceeded 95% of the original sample weight. XANES spectroscopy was performed at the Stanford Synchrotron Radiation Laboratory on beamline 11-2. The storage ring was operated at 3.0 GeV and at currents between 40 and 100 mA. Incident and transmitted intensities were measured with 15-cm, N2-filled ionization chambers. Sample fluorescence was measured with a 13-element Ge detector containing a 6-µm Ge filter. The incident beam intensity was 6824

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detuned approximately 50% to reject higher order harmonic frequencies and to prevent detector saturation. Arsenic K-edge spectra were internally calibrated with sodium arsenate (11 874 eV). XANES spectra were collected by scanning across the K-edge (11 867 eV) using 0.2-eV steps. Spectral processing and data analyses were conducted with the program SixPack (26). The background was removed from the spectrum before normalization using a Gaussian fit for the pre-edge and a quadratic fit for the postedge. As speciation was done by comparing white line features of the model compounds arsenate (as Na2HAsO4), arsenite (as NaAsO2), realgar (AsS), orpiment (As2S3), and arsenopyrite (FeAsS) as shown valid for identifying As oxidation states and various As-sulfide minerals (27–29). To aid in determining the most appropriate standards, first derivatives were analyzed using principle component analysis (PCA) and target transformation. PCA has proven successful in determining statistical variance within an experimental data set composed of unknown samples (30, 31). The number of components necessary to reconstruct the spectra from HP and PP samples was determined from the PCA indicator value, which reaches a minimum between primary components and experimental noise (31). After PCA analysis, spectra were compared to the standards using target transformation. Target transform fits of standards with spoil values less than 2.5 were selected as good fits. Based on these criteria, arsenopyrite, arsenite, and arsenate were selected as standards to quantify As species. Linear combination fitting (LCF) of each unknown spectrum was performed using the base set of standards. LCF was optimized by minimizing reduced χ2 values, and fit results were normalized to unity. Precision of fitting As species from XANES spectra is estimated to be 5% (27).

Results and Discussion Element Abundance. Sediment concentrations of total Mn and Fe vary with depth at both HP and PP, but the trends of such changes differ between the sites (Supporting Information, Figure S1). This apparent inconsistency is a function of differing sedimentation rates at the two sites, with HP receiving on average three to four times the deposition load of PP (3). Sediment cores obtained from sites near those at which equilibrium dialyzers were placed showed that contaminated sediments extended to 119 cm at HP but only 41 cm at PP and that the depth to an identifiable Mt. St. Helens ash layer deposited after the eruption in 1980 was 20.5 to 21.5 cm at HP but only 5.0 to 6.5 cm at PP. Depth-normalized profiles for Mn, Fe, S, and As are shown in Figure 1 to facilitate site comparisons. Changes in Mn concentrations are similar in normalized depth comparisons. Trends in total Fe for the two sites do not follow as closely in absolute concentrations as Mn, but minima and maxima are present near the same respective normalized depths. Total Mn and Fe concentrations at predicted depostion depths of 2 cm are within the range of concentrations measured in suspended sediments collected by Box et al. (2) during the flood event of 1996 (Figure 1A,B). Thus, total elemental concentrations of Mn and Fe reflect a historical record of the material originally deposited. Postdepositional remobilization may modify but does not completely erase the original depositional signature, an observation in agreement with others (3). Total S concentrations increase with depth for both sites (Supporting Information, Figure S1), but depth-normalized profiles for PP sediments contain only about half the total S as HP sediments (Figure 1C), indicating qualitative depositional differences or diagenetic processes that release SO42- to the overlying water. Diagenetic processes are likely responsible for As enrichment at the normalized depths of 1.5 cm in PP sediments as compared to HP (Figure 1, D), in agreement with other reports of surface enrichment of As at

FIGURE 1. A-D. Depth-normalized sediment concentrations of Mn, Fe, S, and As at Harlow Point (HP) and Peaceful Point (PP). Normalized depths were calculated by dividing HP sediment depths by four. Results are averages for total element digests of duplicate cores. The double-sided arrows represent the ranges in respective elemental concentrations for suspended sediments collected in the Coeur d’Alene River (CDAR) during the 1996 flood as reported by Box et al. (2). The arrows are placed at the appropriate depth for materials deposited during the 1996 flood.

FIGURE 2. A, B. Arsenic-XANES derivative spectra (solid lines) of sediments from Harlow Point (A) and Peaceful Point (B) with superimposed linear component fitting (LCF) curves (dotted lines). The vertical dashed lines represent the inflection point of the white line, from left to right, for FeAsS, AsO33-, and AsO43-, respectively. The percentages of FeAsS, AsO33-, and AsO43- were quantified from the LCF results. PP (20, 32) and the observation that As in PP sediments at 2 cm exceeds the range of As concentrations measured in suspended sediments collected by Box et al. (2) during the 1996 flood. Diagenetic changes do not obscure As depositional trends occurring at 4 and 8 cm in the sediment profile (Figure 1D). Organic C in the collected sediments at HP (18.0-27.1 g kg-1) is consistently higher than at PP (13.3-23.1 g kg-1) whether depth corrected for sedimentation or depth-wise comparisons are made (Supporting Information, Figure S2). There was no evidence that greater organic C contributions to HP compressed redox boundaries or influenced changes in the distributions of redox-sensitive elements within sediment porewaters (Supporting Information, Figure S3). Solid-Phase As Speciation. XANES derivative spectra allow identification of predominant As oxidation states in

the solid phase, with some sensitivity for molecular coordination (28, 29). XANES speciation data do not imply that arsenite and arsenate are ionic species, only that these are the representative oxidation states of the species present within the sediment at each respective depth increment. Arsenic in sediments from HP and PP is variable in oxidation state and no clear patterns of species changes with depth are evident at either site (Figure 2). The highest concentrations of arsenate exist when sediment samples contain the lowest concentrations of FeAsS, and in most cases arsenate is present in the lowest percentage with the dominant species being FeAsS and arsenite. High percentages of FeAsS (>68% of the total As) are found at both sites but only at specific depths (HP 6-12 and 30-36 cm; PP 6-12, 12-18, and 30-36 cm) (Figure 2). Recent studies assessing the dissolution of FeAsS have shown that it is stable VOL. 42, NO. 18, 2008 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

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in a variety of environments but have also confirmed that fine-grained FeAsS is reactive in oxic environments (33, 34). Oxidation of FeAsS occurs at circumneutral pH values in the presence of dissolved oxygen concentrations of 0.3 to 17 mg L-1 at a rate of 10-10.14 ( 0.03 mol m-2 s-1 (34), with faster oxidation rates being reported at acidic pHs (35). Although surface passivation may decrease continued oxidation, the oxidized surface layer on arsenopyrite does not offer complete protection to the unoxidized material beneath because it partially dissolves in aqueous solution (36). In addition, the formation of an oxidized surface passivation layer on arsenopyrite provides a reservoir of oxidizing materials that causes continued arsenopyrite oxidation to occur even after the sediments are buried and transitioned to an anaerobic environment (36). Assuming that FeAsS predominately occurs in the form of spherical, silt-sized particles ranging from 2.0 to 50 µm (Supporting Information, Figure S4) having a density of 6.0 g cm3 (37), oxidation rates vary from 1% a year for 50-µm FeAsS particles, up to 20% a year for 2.0-µm particles. These calculations indicate that oxidation is probable when FeAsS remains in the top 5 cm of sediment. Geochemical predictions and field observations do not support in situ precipitation of FeAsS (14, 27, 33), making it likely that FeAsS in Lake CDA sediments is detrital and not authigenic. Arsenopyrite is found within mineral deposits of the region (38), and its presence in Lake Coeur d’Alene sediments has previously been confirmed with SEM (3, 20). Although not identified in our samples, other As sulfides present in mineral deposits of the region and potentially present in Lake Coeur d’Alene sediments include tetrahedrite and gersdorffite (38). Due to the fine-grained nature of these sediments and the reactivity of fine-grained FeAsS, preservation of this mineral is therefore a consequence of rapid transition by burial through the oxic environment of the surface sediments. The variable concentration of FeAsS in the sediments, 0 to 86% of the total As (Figure 2), implies that the deposited material had widely different FeAsS concentrations or that the extent of FeAsS oxidation varied throughout history. Total elemental (Figure 1 and Supporting Information, Figure S1) and XANES analyses (Figure 2) indicate that deposited materials varied temporally, but that FeAsS residence time in the oxic sediment zone may also induce diagenetic alteration. Flooding events in the CDA River Basin are frequent, mobilizing material from the riverbanks and within the river channel, subsequently depositing this material throughout the floodplain and in Lake CDA. There are compositional differences in the sediment loads of these flood events, because larger events erode material from different sources than smaller flood events. Riverbed sediments are typically scoured, mobilized, and transported during the lesser flood events, whereas suspended sediments mobilized during larger flood events are dominated by materials derived from floodplain soils (2). Although we cannot definitively identify the sources for suspended sediment mobilized during the larger flood events, deposited materials possess distinct geochemical signatures that are partially responsible for the observed temporal variation in As-XANES spectra. Examples of differing sediment sources and loads are reflected in the XANES spectra for flood events occurring in February 1996 and January 1974. The February 1996 flood was the second largest event recorded since 1911. This flood produced suspended-sediment loads that were 2 orders of magnitude greater than three other floods from 1995 to 1997 (2), depositing sediment now found in HP cores at 6-12 cm and PP cores at 0-3 cm. This large flood event eroded soils or tailings having high FeAsS concentrations, thus yielding an As-XANES spectrum for HP showing 68% FeAsS (Figure 2A). Depositional events represented in the 6-12-cm sample 6826

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at HP had sufficient sediment volume to transition FeAsS from the oxic zone to the suboxic zone where the mineralogical signature was preserved. Since 1996, lower river flows eroded highly weathered riverbed and stream bank sediments producing the spectra observed for HP samples obtained from 0-3 and 3-6 cm. These spectra indicate that only 10-27% of the As is present in the form of FeAsS (Figure 2A). High FeAsS percentages are not observed at 0-3 cm in PP sediment samples not only because of greater FeAsS oxidation but also because sediment with lower percentages of FeAsS diluted the signature of material deposited by the 1996 flood. Preferential sedimentation of high density FeAsS particles at HP would also serve to enhance FeAsS concentration differences between the sites. Thus, although there is a shoulder in the spectrum indicating FeAsS in the 0-3 sample (Figure 2B), LCF of the XANES spectrum excluded it as a dominant component. A second flood event that provides additional evidence for differences in FeAsS loading is the largest event ever recorded. The magnitude of this 1974 flood produced signatures similar to the 1996 event not only because the sediment included materials with higher FeAsS content but also because deposited sulfidic minerals rapidly transitioned through the oxic zone and into the suboxic zone where they retained their sulfidic signature. As a result of depositional differences, the 30-36-cm sample at HP corresponds to the 6-12-cm sample obtained from PP, with both having similar total As concentrations (Figure 1D and Supporting Information, Figure 1S, G,H). The As-XANES spectra from these two depths are similar with 84 to 86% of the total As in the form of FeAsS (Figure 2A,B), confirming that the sediments are likely of the same origin and that little postdepositional alteration has occurred. The magnitude of this flood event was large enough to preserve the FeAsS signature even within the lower depositional environment of PP. Porewater Arsenic. Arsenic appears in the porewater of HP at 5 cm, showing a continuous increase to approximately 25 cm where the maximum concentration of 8.4 µM was detected (Figure 3). The trend at PP is different in that after the first detection of As in the porewater at 5 cm, the maximum concentration of 16.2 µM was measured within the next few centimeters (Figure 3). Once this maximum is established, the remainder of the profile shows some variation, but the aqueous concentrations of As always exceeded those found at HP. Porewater concentrations of As at HP are 367% and PP 768% of the CCC of 2.0 µM adopted for the protection of freshwater biota (39), thus creating a significant hazard to this ecosystem (7, 40). Even with these high concentrations of As in the porewater, As falls to less than 0.01 µM within the upper 3 cm of sediments and in the overlying water at both sites (Figure 3). We propose the sequence shown in Figure 4 to explain observed changes in sediment and porewater As in Lake CDA. Eroded materials highly variable in composition are deposited to the sediment-water interface where oxidative reactions facilitate the dissolution of arsenic sulfides and the formation of Fe-oxhydroxides. The oxidation of FeAsS results in the concomitant release of As and SO42- to the aqueous phase. Soluble As binds to the Fe-oxhydroxides, whereas a fraction of SO42- remains within the porewater and diffuses into the overlying water column. The simultaneous decline of Fe and As in porewaters obtained from the sediment-water interface (Supporting Information, Figure S3 and Figure 3) indicates sorption of soluble As whether originating from oxidative dissolution of FeAsS or diffusion from suboxic/anoxic sediment depths. With additional deposition, As sorbed at the oxic sediment-water interface transitions to a suboxic/anoxic region where reductive processes release soluble As to the

FIGURE 3. Aqueous concentrations of As in the sediments and overlying water from duplicate dialyzers placed at Harlow Point and Peaceful Point. Data points at depths less than zero represent the overlying water.

FIGURE 4. Processes controlling As cycling in freshwater sediments where periodic events deposit additional contaminated materials, transitioning oxic sediments to a suboxic/anoxic environment. Precipitation (open arrow), dissolution (dotted arrow), and sorption (dashed arrow) are shown. porewater (Figure 4). The dissolution of Fe-oxyhydroxides as indicated by high concentrations of dissolved Fe (Supporting Information, Figure S3), correlates with increases in dissolved As in the sediment porewater (Figure 3), consistent with predictions of others (15, 18, 41–43). Recent investigations of As speciation in sediments of smaller contaminated lakes within the Coeur d’Alene Mining District have shown that unstable redox conditions as controlled by fluctuating water levels similarly increase As concentrations in the sediment pore waters (44). In the anoxic zone, we expect SO42- reduction and precipitation of FeSx or other metal sulfides (13). Sulfidic minerals typically sequester As from solution by either adsorption or coprecipitation (45), thereby decreasing dissolved concentrations. However, the loss of SO42- during sediment exposure to oxic conditions at the sediment-water interface in combination with the high Fe:S ratios of sediments limit the amount of sulfide available for As precipitation (19). A similar efflux of SO42- leading to increased soluble As also occurs in As-contaminated sediments of Bangladesh (46). As a result of limited sulfide precipitation, porewater As remains at biologically harmful concentrations (Figure 3). Although sedimentation differences have been shown to affect porewater concentrations of As in a marine environment (43), we are unaware of any freshwater illustrations of such phenomena. Larger or more frequent deposition events as occur at HP decrease the time available for oxidative diagenetic reactions by burying the surface material, repo-

sitioning it to a more reduced zone. In contrast, less sediment deposition and longer sediment residence times at PP lead to increased FeAsS oxidation and correspondingly more As accumulation on iron oxides at the oxic sediment-water interface. Our data (Figure 1D and Supporting Information, Figure S1) as well as that of others (20) indicate greater surface enrichment of As in PP than HP sediments. The trend for lower total S concentrations at PP is also consistent with greater dissolution of sulfides in the oxic zone and diffusional losses of SO42--S into the overlying water (Figure 1C). Greater amounts of As accumulation and less sediment deposition at PP ensure that more soluble As will be released at shallower sediment depths in PP than HP sediments, yielding the observed porewater As differences (Figure 3). Because more sulfate was lost from PP sediments prior to burial, less sulfide is available for As precipitation and porewater As concentrations increase to a greater concentration. This cycle is further enhanced by high porewater As concentrations just below the oxic sediment cap in PP sediments, thereby establishing a diffusional gradient that promotes additional As accumulation on iron oxides in the overlying oxic zone. A two-step process in which As trapped in the oxic zone is released in soluble form in an anoxic, S-limited environment virtually ensures that soluble As in the sediment porewaters of Lake CDA will remain at toxic levels (Figure 4). This process presents a dilemma for lake management since aerobic conditions at the sediment-water interface trap As, preventing its flux and subsequent dilution in the VOL. 42, NO. 18, 2008 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

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overlying water; however, burial and reductive dissolution of the trapped As produces toxic As concentrations in the anoxic sediment porewater. Thus due to redox stratification in the sediment and periodic depositional events, a decrease in dissolved As concentrations within the sediments will only occur by minimizing As inputs to Lake CDA.

Acknowledgments Research was supported by funding from the Idaho Water Resources Research Institute, the Inland Northwest Research Alliance, EPA-STAR Fellowship Program, and the Stanford NSF-Environmental Molecular Science Institute. Portions of this research were performed at the Stanford Synchrotron Radiation Laboratory (operated by Stanford University), a national user facility supported by the U.S. DOE, Office of Basic Energy Sciences with additional funding from NIH. We thank C. Cornwall and J. Failla for their expertise and L. Balistrieri for advice.

Supporting Information Available Figures showing total sediment metal(loid) and organic carbon concentrations, porewater elements, and sediment particle size. This information is available free of charge via the Internet at http://pubs.acs.org.

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