Environ. Sci. Technol. 2000, 34, 3131-3136
Microbial Reduction of Arsenate in the Presence of Ferrihydrite HEIKO W. LANGNER* AND WILLIAM P. INSKEEP Department of Land Resources and Environmental Sciences, Montana State UniversitysBozeman, Bozeman, Montana 59717-3120
Increased mobilization of As under anaerobic conditions is of great concern in As contaminated soils and sediments. The identification of important release mechanisms may assist in designing safe and effective remediation strategies. In this study we investigated the effect of microbial reduction of aqueous arsenate (As(V)) on the solubilization of As(V) sorbed to ferrihydrite, in the absence of reductive dissolution of the Fe(III)-oxide solid phase. The addition of 0.1, 1.0, and 5.0 mM As(V) to serum bottles containing 10 mmol L-1 Fe(III) as ferrihydrite resulted in the sorption of 98, 75, and 20% of the applied As(V), respectively. Inoculation with an As(V) reducing, glucose fermenting microorganism (CN8) was followed by complete reduction of aqueous As(V) to As(III) at nontoxic As concentrations (up to 1.0 mM), but no reduction or dissolution of the Fe(III) solid phase was observed. Despite rapid reduction of aqueous As(V) to As(III), sorbed phase As remained primarily as As(V), and desorption of As(V) was too slow to cause a significant increase in aqueous As concentration over the 24-day experiment. Our study suggests that the reduction of aqueous As(V) may play a relatively minor role in the solubilization of As(V) sorbed to Fe (hydr)oxide. Arsenic release from contaminated soils and sediments may proceed considerably faster under conditions favoring dissimilatory Fe(III) reduction leading to the dissolution of sorbing phases.
Introduction Elevated concentrations of arsenic (As) in natural waters associated with mine tailings, geothermal areas, and As-rich parent materials can represent a major water quality and health problem for humans (1-3). The environmental and toxicological impact of As is largely determined by the distribution of the element between immobile and mobile forms and by the distribution of valence states, primarily arsenate (As(V)) and arsenite (As(III)) (4, 5). Iron (hydr)oxides are considered important solid phases responsible for As sorption and immobilization in aerobic soils and sediments (6). However, significant solubilization of sorbed As is commonly observed in reducing environments where As(III) becomes the predominant aqueous As species (7, 8). Elucidation of important pathways responsible for As(V) reduction and increases in total aqueous As would be beneficial for designing watershed management practices that minimize surface or groundwater contamination. Mechanisms responsible for increases in total aqueous As under anaerobic conditions may vary considerably * Corresponding author phone: (406)994-5578; fax: (406)994-3933; e-mail:
[email protected]. 10.1021/es991414z CCC: $19.00 Published on Web 06/16/2000
2000 American Chemical Society
depending on local environmental conditions. Certainly, microbially mediated reductive dissolution of Fe(III) solid phases is an important process contributing to the release of phosphate and trace metals from soils and sediments (911). Jones et al. (12) used a mixed microbial population from a naturally contaminated As soil to study As release from goethite and ferrihydrite during microbial growth. Reductive dissolution of the Fe oxide phase was an important mechanism of As release primarily for ferrihydrite; the lower surface area goethite sample did not support significant rates of reductive dissolution. This is consistent with relationships established by Roden and Zachara (13) using the Fe(III) reducing bacteria Shewanella alga strain BrY, showing greater reductive dissolution rates for high surface area ferrihydrite than well crystalline goethite or hematite. Another recent study (14) demonstrated the liberation of As(V) from scorodite (FeAsO4) and lake sediments by strain BrY. Simultaneous As(V) and Fe(II) accumulation in the aqueous phase indicated that BrY was able to couple respiration with reduction of Fe(III) resulting in the dissolution of scorodite. Although As(V) was subsequently solubilized, BrY was not capable of reducing As(V) under their experimental conditions. In natural environments containing mixed microbial populations, liberated As(V) would likely be reduced and result in an accumulation of As(III) instead of As(V). Collectively, these studies emphasize the potential role of microorganisms in solubilizing As associated with Fe(III) solid phases, independent of whether As(V) is reduced to As(III). Increases in total aqueous As concentrations observed under reducing conditions may also be due to desorption of As(V) in response to microbial reduction of aqueous phase As(V) to As(III). At circumneutral pH values and low As:Fe ratios present in most environments, As(III) sorbs less strongly to Fe(III) (hydr)oxides (15); consequently, transformation of aqueous As(V) to As(III) may initiate reequilibration of sorption reactions resulting in desorption of As(V). In environments where aqueous As(V) is rapidly reduced (time scale of hours), the rate of As(V) desorption may become the rate-determining step controlling solubilization of sorbed As(V). This would be especially important in systems where rates of reductive dissolution are low, as is the case for lowsurface area Fe(III) hydroxides (12, 13). The varied microbial processes important in As cycling underscore the need for continued inquiry regarding As transformations in water, soil, and sediments, in hopes of developing greater predictability of As behavior in landscape settings. We also recognize that other As transformation pathways may make very important contributions to the fate of As. For example, precipitation-dissolution reactions of As(III)-sulfide phases in sulfate-reducing environments and abiotic surfacecatalyzed oxidation-reduction reactions are important processes (16-18) but are not the focus of the current study. The microorganism used in the current study is a glucosefermenting, Clostridium-like isolate enriched from an Ascontaminated soil (12). This isolate was shown to rapidly (1 d) reduce aqueous As(V) under fermentative conditions but not as a result of coupling As(V) reduction with respiration (12). The mechanism of reduction may be consistent with As-detoxification pathways described for a number of bacterial isolates (19, 20). Consequently, this study was designed to determine the importance of microbially mediated As(V) reduction on solubilization of As(V) sorbed to ferrihydrite, in the absence of significant reductive dissolution of the Fe(III)-solid phase. VOL. 34, NO. 15, 2000 / ENVIRONMENTAL SCIENCE & TECHNOLOGY
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Materials and Methods The reduction of As(V) was studied in the presence and absence of ferrihydrite using a glucose-fermenting Clostridium sp. (CN8) enriched from an As-contaminated soil (12). All microbial growth experiments were conducted using closed-headspace serum bottles (72 mL), filled initially with 50 mL of minimal salts solution containing 6 mM NH4Cl, 2 mM MgCl2, 1 mM CaSO4, 50 µM KH2PO4, 2.5 mM NaOH, 5 µM FeCl2, 100 µL L-1 of micronutrient solution (21), and 0.2 g L-1 Sigma Select yeast extract (6.5 mM C). Glucose (30 mM C) was used as a C source for microbial growth; nonradiolabeled and uniformly 14C-radiolabeled (Nycomed Amersham plc, Buckinghamshire, U.K.) glucose were mixed to yield specific activities of 5 × 106 Bq (mol C)-1 or 5 × 107 Bq (mol C)-1 in treatments selected for ion chromatographic/flow scintillation analysis. A subset of serum bottles contained 50 mg of 2-line ferrihydrite (amorphous Fe oxide (22); 298 ( 13 m2 g-1 surface area as measured by N2 adsorption at 77 K) representing a concentration of 10 mmol L-1 solid-phase Fe(III). Aliquots of Na2HAs04‚7H2O stock solutions were added to yield total As(V) concentrations of 0, 0.1, 0.3, 1.0, or 5.0 mM. The solution also contained 20 mM KHCO3 and 0.15 mL of 6 M HCl to buffer pH at 6.8 ( 0.2. Capped serum bottles were autoclaved and equilibrated for 4 d, standing upright on a rotary shaker (120 min-1) at 23 °C in the dark. Nonsterile treatments were then inoculated with 0.15 mL of a cell suspension containing approximately 5 × 108 cells mL-1 (DAPI staining method) and incubated under the same conditions. Serum bottles were sampled (10 mL) periodically over 24 d under N2(g) atmosphere and aseptic conditions through the rubber septa and analyzed for C, As, S, and Fe species. Three separate 1-mL subsamples were analyzed for 14C using liquid scintillation (Model 2200CA Liquid Scintillation Analyzer, Packard Instrument Co.) as follows: (i) an unfiltered untreated sample, (ii) an unfiltered, acidified (0.01 mL 12 M HCl) and swirled (30 min) sample, and (iii) a filtered (0.22 µm), acidified, and swirled sample. Estimates of aqueous 14C-carbonate content were obtained by subtracting the 14C content of the unfiltered acidified samples from the 14C content of the unfiltered untreated samples. The amount of glucose-C incorporated into biomass was estimated from the difference in 14C-contents between unfiltered acidified and filtered acidified samples. Aqueous metabolites of glucose fermentation were identified in selected serum bottles using gas chromatography-mass spectroscopy (GC-MS, Model 5890 Series II+ GC [Hewlett Packard Co.] equipped with a Stabilwax [30 m length/0.32 mm ID/1 µm drum thickness, Restek Corp.] column; Model VG70E double focusing mass spectrometer [Micromass Ltd., Manchester, U.K.]). Organic acids were quantified using ion chromatography (IC) with suppressed electrical conductivity detection (Dionex AS6ICE column; mobile phase 80% 0.4 mM aqueous heptafluorobutyric acid and 20% acetonitrile; AMMS-2 suppressor). Glucose concentrations were analyzed simultaneously by directing the IC column eluant through a 14C-radioisotope detector (IC-RID; Packard Instr., Radiomatic 500TR Series flow through scintillation analyzer). Samples for the determination of aqueous As were split immediately for separate analysis of total aqueous As and As(V). For As(V), 1 mL of filtered sample was added to 4 mL of H2O and 1 mL of 2 M TRIS buffer (pH 6.0). Dilution with H2O was necessary to minimize the amount of sample used and to avoid interferences with aqueous Fe during the removal of As(III). While sparging the mixture with N2(g), 1 mL of a solution containing 0.25 M NaOH and 0.79 M NaBH4 was added in 0.2-mL increments over 4 min to reduce As(III) to arsine gas. The mixture was N2(g) sparged for an additional 3132
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3 min to purge arsine. This sample (containing As(V) only) and the original sample (containing total As) were preserved in 0.1 M HCl prior to As analysis using hydride generation atomic absorption spectrophotometry (HG-AAS (12)) or inductively coupled plasma emission spectrometry (ICP), for samples containing As concentrations > 5 µM. Although our instrumental detection limit (3 SD of 50 absorbance blank readings) for aqueous As using HG-AAS was 3.4 nM, necessary dilutions and variable sample composition allowed us to reliably quantify As concentrations above 0.5 µM. As(III) concentrations were determined as the difference between concentrations of total aqueous As (original sample) and As(V). Total aqueous S concentrations were determined with ICP, and selected samples were analyzed for sulfate using IC (Dionex AS4A-SC column) and sulfide using the methylene blue colorimetric method (23). Aqueous Fe(II) and Fe(III) concentrations were determined colorimetrically using phenanthroline (23). To determine Fe(II) and As species bound by the ferrihydrite solid phase, serum bottles were sampled destructively. For Fe(II), aqueous Fe(II) was determined first in a 2.5-mL sample, followed by addition of 12 M HCl to yield 0.5 M HCl. Bottles were shaken for 1.5 h, and Fe(II) was determined colorimetrically. The solid-phase Fe(II) concentration was then calculated by difference between Fe(II) after and before the HCl extraction. Solid-phase As was speciated in a different set of serum bottles using a method modified from Sun and Doner (24). After sampling the aqueous phase, 10 M NaOH was injected anaerobically through the rubber septa to yield 1 M NaOH. Serum bottles were then shaken for 2 h in a water bath at 70 °C, carefully depressurized (syringe needle through septum), decapped, and sampled. Filtered (0.22 µm nylon filters) 1-mL subsamples were added to 1 mL of 1.0 M HCl and 3 mL of H2O, and As speciation and analysis were conducted as described for aqueous samples. Solid-phase As(V) and As(III) concentrations were calculated as the difference between As(V) and As(III) before and after extraction with NaOH. The method described above for analyzing sorbed As species on ferrihydrite was selected after testing several extractants including phosphate (0.1 M K2HPO4/KH2PO4, pH 7.0, 5 min), HCl (0.5 M, 1.5 h), and NaOH (1.0 M, 2 h, 70 °C) under aerobic conditions in the dark at various sorbed As concentrations and As(III):As(V) ratios. Each extractant was tested using nine treatments, each in duplicate. Arsenic was sorbed to ferrihydrite at As:Fe ratios of 0.1:10 (mmol L-1) and 1.0:10 (mmol L-1) each at As(V):As(III) ratios of 0:1, 0.3:0.7, 0.7:0.3, and 1:0. After 4 d of equilibration, ferrihydrite with sorbed As was washed with background solution followed by quantification of unsorbed As species. Washed solid phases were then immediately exposed to the extractant. One additional treatment included no ferrihydrite and a 0.05: 0.05 mM As(V):As(III) mixture to examine the effect of the extractant on the aqueous redox species. The NaOH extraction method outperformed other methods and, under aerobic conditions, yielded recoveries of 130% (standard error of 7%, N ) 6) for As(V) and 38(2)% for As(III) at the high As:Fe ratio and 91(1)% for As(V) and 35(3)% for As(III) at the low As:Fe ratio (Table 1). As(V) recoveries greater than 100% suggested partial oxidation of As(III) during the extraction procedure and were also observed in the control treatment without ferrihydrite sorbent (recoveries of 128 ( 0% for As(V) and 60 ( 5% for As(III)). Consequently, the extraction was repeated under anaerobic conditions (70 °C; 2 h; N2-purged solution, headspace, and NaOH) yielding recoveries of 48(5)% As(V) and 95(18)% As(III) at the high As:Fe ratio and 78(11)% As(V) and 53(16)% As(III) at the low As:Fe ratio. Control treatments without the addition of solid phase and exposed to the anaerobic extraction procedure yielded recoveries of 86(6)%
TABLE 1. Recoveries of Sorbed As(V) and As(III) from Ferrihydrite Using Various Extractants recoverya (%) As:Fe ) 1:10
As:Fe ) 1:100
extractant
As(V)
As(III)
As(V)
As(III)
1 M NaOH (70 °C) 0.5 M HCl 0.1 M PO43- (pH 7) NaOH (anaerobic)
130(7)b 16(2) 4(1) 48(5)
38(2) 36(6) 3(1) 95(18)
91(1) 15(2) 8(1) 78(11)
35(3) 44(8) 2(1) 53(16)
a Recoveries are averages over all As(V):As(III) ratios with each species representing 100, 70, and 30% of the total applied As. Standard errors in parentheses (N ) 6). b Recovery > 100% for As(V) suggests oxidation of As(III) under aerobic conditions.
FIGURE 1. Reduction of aqueous phase As(V) in the absence of ferrihydrite after inoculation with CN8, a glucose-fermenting Clostridium sp. (12). Symbols represent averages, and error bars show standard deviations of samples from 12 replicates. As(V) and 99(10)% As(III) indicating the stability of As redox states during extraction. Anaerobic conditions generally improved As(III) recoveries and were therefore used in the analysis of experimental samples. Large standard errors were due primarily to systematic changes in recoveries as a function of As(III):As(V) ratios, rather than variation among replicates. For example, the As(III) recovery (anaerobic experiment, low As:Fe ratio) was 107 ( 12% when As(III): As(V) ) 0.3:0.7 but decreased to a minimum of 19 ( 8% when exclusively As(III) was present on the sorbent. This trend suggests that As(III) desorbs more readily in the presence of As(V) which is consistent with recent findings (25). More importantly, the extraction tests with NaOH indicated that the method was suitable to reliably detect sorbed phase As(III) at As(III):As(V) ratios as low as 0.3:0.7.
Results and Discussion Arsenate Reduction in the Absence of Ferrihydrite. A series of As(V) reduction experiments was conducted in the absence of ferrihydrite in serum bottles containing minimal salts media, 30 mM C (as glucose) and 0.3 mM As(V) at the time of inoculation with the Clostridium sp., strain CN8. Sixty percent of As(V) was reduced to As(III) within 1 d, corresponding to the period of maximum microbial growth rate (Figure 1). This is consistent with results of Jones et al. (12) showing a first-order dependence of As(V) reduction rate on both As(V) concentration and microbial biomass during the microbial growth phase and at nontoxic As levels (e0.6 mM As). Growth rates were not affected by the presence of As(V), based on a comparison of optical densities in serum bottles with and without As(V). In fact, rates of conversion of glucose14C to biomass, carbonate, and fermentation products were not substantially different in any treatments of this study. The sole deviation occurred in the treatment containing 5.0 mM total initial As(V) (about 4 mM dissolved arsenate), where
FIGURE 2. Fate of applied 14C-glucose in all inoculated treatments with and without ferrihydrite at total (sorbed + aqueous) As concentrations between 0 and 1.0 mM (panel A) and in treatments with a total As concentration of 5 mM (panel B). Values are averages based on 14C analysis. Error bars represent standard deviations about the mean of all treatments included; missing error bars indicate that no replicate samples were taken. The composition of dissolved organic 14C changed during microbial growth from glucose to a mixture of fermentation products including butyrate (butanoate) as the primary compound. As toxicity caused a delay in microbial growth of approximately 1 d (Figure 2A,B). Disappearance of glucose was accompanied by the appearance of organic fermentation products (butyrate (butanoate), acetate, formate and derivatives) and H2(g), identifying CN8 as a glucose-fermenter (12). This suggests that As(V) reduction by CN8 may be part of a detoxification mechanism, rather than dissimilation (19). Sulfate was not reduced by CN8, and its aqueous concentration remained essentially unchanged at 0.77 ( 0.05 mM throughout all treatments (data not shown). Reduction of Aqueous As(V) in the Presence of Ferrihydrite. The ability of CN8 to reduce aqueous versus sorbed As(V) was studied in the presence of ferrihydrite (10 mM Fe) at three total As(V) levels, which resulted in different ratios of sorbed:aqueous As(V). At an initial total As(V) concentration of 0.1 mM, less than 2% (2 µM) of the As(V) was present in the aqueous phase (Figure 3A); the remaining As(V) was sorbed to ferrihydrite (Figure 3D). Total aqueous As decreased in both inoculated and sterile treatments to about 0.8 µM during the first 10 d of the experiment suggesting a slow sorption process not affected by microbial activity. In the sterile treatments, As(V) was the predominant form of As throughout the experiment. However, aqueous As(V) concentrations in the inoculated treatments dropped below 0.5 µM within the first day, indicating reduction of aqueous phase As(V) to As(III). Although total aqueous As concentrations were just above our analytical detection levels for the 0.1 mM As treatment, it is clear that CN8 reduced aqueous As(V) to As(III) but did not trigger significant solubilization of As sorbed to the Fe oxide phase. At an initial total As(V) concentration of 1.0 mM, approximately 25% of the As (0.25 mM) was aqueous (Figure 3B), while 75% was sorbed to the ferrihydrite phase (Figure 3E). Aqueous As(V) was nearly completely reduced to As(III) within 1 d after inoculation with CN8. Similar to the 0.1 mM As treatment, growth of VOL. 34, NO. 15, 2000 / ENVIRONMENTAL SCIENCE & TECHNOLOGY
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FIGURE 3. Speciation of aqueous As (panels A-C) and sorbed phase As (panels D-F) at total (aqueous + sorbed) As concentrations of 0.1, 1.0, and 5.0 mM, where 99, 75, and 30%, respectively, of the total As were sorbed to 10 mmol L-1 Fe(III) oxide (as ferrihydrite). The difference between total aqueous As and aqueous As(V) concentrations represents the concentration of aqueous As(III). Error bars show standard deviations about the mean of all replicate treatments sampled at a specific time. CN8 did not result in solubilization of As from the Fe oxide phase. In treatments containing 5.0 mM total As, approximately 80% (4 mM) remained in the aqueous phase (Figure 3C). These higher levels of soluble As(V) caused a delay in both the growth of CN8 and the reduction of As(V) by approximately 1 day compared to treatments with lower aqueous As(V) concentrations. As(V) reduction was detectable concurrent with the onset of microbial growth, which was first observed in the samples taken at 1.7 d (Figure 3C). The absolute mass of As(V) reduced to As(III) during the growth phase was greater than in treatments at lower As(V) concentrations. However, only 1 mM As(V) was reduced to As(III) leaving approximately 3 mM As(V) in solution until termination of the experiment (Figure 3C). Oxidation State of Sorbed Phase As. Assuming equilibrium and reversible sorption conditions for As(V) and As(III), one might expect that the reduction of aqueous As(V) to As(III) would result in adsorption of As(III) and desorption of As(V). Assuming further that the sorption coefficients for As(III) and As(V) are different, this should have resulted in a net decrease or increase in the aqueous As fraction compared to the sterile controls where no As(V) reduction occurred. However, observed changes in aqueous As concentrations were insignificant with respect to total concentrations of As in the systems. For example, aqueous As concentrations in the 0.1 mM As treatment changed from 1.8 µM to 0.8 µM (Figure 3A, inoculated treatment) representing a phase change for only 1% of the total As in the system. Similarly, at total As levels of 1.0 and 5.0 mM, the distribution of As between the phases remained essentially constant over the 24-d experiment, despite significant reduction of aqueous As(V) to As(III). Further, results from extraction and speciation of sorbed As showed no evidence of appreciable amounts of sorbed As(III) (Figure 3D-F); As(V) was the predominant form of sorbed As throughout the experiments. Mass balance calculations showed that the sum of aqueous As plus sorbed phase As (add As levels in Figure 3134
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3 (panels A and D, B and E, or C and F)) was in agreement with expected total levels of As at 0.1, 1.0, and 5.0 mM, confirming that the NaOH extraction procedure adequately recovered sorbed As from the ferrihydrite phase. Consequently, several lines of evidence suggest that (i) aqueous As(III) formed from the reduction of As(V) was not resorbed by the ferrihydrite phase, and (ii) the reduction of aqueous As(V) to As(III) did not encourage significant desorption of As(V). Relative to rates of microbially mediated reduction of aqueous As(V), it appears that rates of sorption and desorption processes involving As species were slow on the aged and possibly biofilm-coated ferrihydrite. Stability of Fe(III) Oxide Sorbent. The stability of the ferrihydrite phase during growth of CN8 was also confirmed by analysis of Fe(II) in aqueous samples and Fe oxide samples extracted with 0.5 M HCl (13), in both sterile controls and in vessels containing CN8. While Fe(II) was detectable in both aqueous and solid phases, no substantial increases in total Fe(II) were noted during growth of CN8 (Figure 4). A slightly higher Fe(II) solubility was observed in all inoculated treatments as compared to sterile controls which was consistent with a decrease in pH by 0.29 ( 0.02 units during glucose fermentation. In contrast, microbial reduction of Fe(III) reported by Roden and Zachara (13) resulted in considerably higher yields of Fe(II) (0.28 nmol min-1 m-2 for nonaggregated Fe(III) oxides) than in our treatments (0.04, 0.03, or 0.00 nmol min-1 m-2 in our 0.1, 1.0, or 5.0 mM As treatments, respectively). It is evident from these results that significant reduction of solid phase Fe(III) did not occur during growth of CN8, consistent with observations that CN8 was fermenting glucose, and was not using Fe(III) as a terminal electron acceptor coupled to anaerobic respiration. Although the Fe(III) oxide phase is not thermodynamically stable relative to Fe(II) under anaerobic conditions, abiotic reduction rates were inconsequential during the course of these experiments. Furthermore, constant levels of Fe(II) in the presence of As(III) as observed in our inoculated
reduction of As(V) to As(III) may be occurring via pathways other than dissimilation such as detoxification mechanisms reported in several bacterial genera (19-20). The reduction of As(V) by the Clostridium sp. used in our study appears to occur via a similar mechanism (12). In natural systems, it is expected that a host of abiotic and biotic processes contribute to solubilization and reduction of As from contaminated sediments under reduced conditions. The current study provides an example of one of the potential contributing processes to As cycling: rapid microbial reduction of aqueous As(V) in the absence of significant dissolution of the Fe oxide phase. Under these conditions, sorbed As remains predominantly As(V) and is not solubilized during microbial growth.
Acknowledgments Although the research described in this article was funded in part by the United States Environmental Protection Agency through a grant (R825403-01-0) to W. Inskeep, it has not been subjected to the Agency’s required peer and policy review and therefore does not necessarily reflect the views of the Agency, and no official endorsement should be inferred. This work was also supported with funds from the Montana Agricultural Experiment Station (913300).
Literature Cited
FIGURE 4. Average Fe(II) concentrations in aqueous and solid phases of all treatments containing ferrihydrite (10 mM total Fe) and initial As(V) concentrations between 0 and 1.0 mM. Dissolved Fe(III) concentrations followed thermodynamic predictions in that they were below detection. The slightly elevated aqueous Fe(II) concentration in the inoculated treatments is consistent with higher Fe(II) solubility upon decrease in pH during glucose fermentation. treatments did not suggest any oxidation of As(III) by Fe(III), which had been reported by de Vitre et al. (26) at a molar As:Fe ratio of 1:3400. Our results are consistent with other reports at As:Fe ratios more similar to this study (27) where Fe(III) did not contribute to the oxidation of As(III). Results of our study suggest that the reduction of aqueous As(V) may play a relatively minor role in the solubilization of As(V) sorbed to Fe (hydr)oxide. While aqueous As(V) was reduced by CN8 within 1 d, desorption rates of As(V) from ferrihydrite were too slow to cause a significant increase in aqueous As concentrations over the 24-d duration of the experiment. Arsenic release from contaminated soils and sediments may proceed considerably faster under conditions favoring dissimilatory Fe(III) reduction and subsequent dissolution of Fe(III) sorbing phases. Certainly, many microorganisms are capable of utilizing Fe(III) as an electron acceptor during respiration (9, 12, 18). In such cases, rates of reductive dissolution of the Fe oxide phase would likely exceed rates of As(V) desorption. Several studies have demonstrated that certain heterotrophic microorganisms can couple As(V) reduction with oxidation of lactate (2-hydroxypropanoate) and acetate (28), potentially explaining the observation that As(III) is often the predominant valence state in reduced environments (4, 5). However, several studies have reported the presence of As(III) in environments where dissimilatory reduction of As(V) would not be expected (2931). Consequently, in many soil and water environments,
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Received for review December 19, 1999. Revised manuscript received May 1, 2000. Accepted May 1, 2000.
ES991414Z
