Geochemical and Microbiological Responses to Oxidant Introduction

Jul 11, 2014 - ABSTRACT: Pliocene-aged reduced lacustrine sediment from below a subsurface redox transition zone at the 300 Area of the Hanford site ...
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Geochemical and Microbiological Responses to Oxidant Introduction into Reduced Subsurface Sediment from the Hanford 300 Area, Washington Elizabeth M. Percak-Dennett and Eric E. Roden* Department of Geoscience, University of WisconsinMadison, 1215 West Dayton Street, Madison, Wisconsin 53706, United States S Supporting Information *

ABSTRACT: Pliocene-aged reduced lacustrine sediment from below a subsurface redox transition zone at the 300 Area of the Hanford site (southeastern Washington) was used in a study of the geochemical response to introduction of oxygen or nitrate in the presence or absence of microbial activity. The sediments contained large quantities of reduced Fe in the form of Fe(II)-bearing phyllosilicates, together with smaller quantities of siderite and pyrite. A loss of ca. 50% of 0.5 M HCl-extractable Fe(II) [5−10 mmol Fe(II) L−1] and detectable generation of sulfate (ca. 0.2 mM, equivalent to 10% of the reduced inorganic sulfur pool) occurred in sterile aerobic reactors. In contrast, no systematic loss of Fe(II) or production of sulfate was observed in any of the other oxidant-amended sediment suspensions. Detectable Fe(II) accumulation and sulfate consumption occurred in non-sterile oxidant-free reactors. Together, these results indicate the potential for heterotrophic carbon metabolism in the reduced sediments, consistent with the proliferation of known heterotrophic taxa (e.g., Pseudomonadaceae, Burkholderiaceae, and Clostridiaceae) inferred from 16S rRNA gene pyrosequencing. Microbial carbon oxidation by heterotrophic communities is likely to play an important role in maintaining the redox boundary in situ, i.e., by modulating the impact of downward oxidant transport on Fe/S redox speciation. Diffusion−reaction simulations of oxygen and nitrate consumption coupled to solid-phase organic carbon oxidation indicate that heterotrophic consumption of oxidants could maintain the redox boundary at its current position over millennial time scales.

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INTRODUCTION Subsurface redox cycling of iron (Fe) and sulfur (S) can exert key controls on the speciation and transport of redox-sensitive metals and radionuclides, such as U and Tc. Fe(III)-bearing phyllosilicates, Fe(III) oxides, and sulfate can serve as electron acceptors for dissimilatory Fe(III)- or sulfate-reducing bacterial activity, potentially resulting in enzymatic U(VI) or Tc(VI) reduction and immobilization.1 Additionally, biogenic reduced Fe/S-bearing phases can serve as abiotic reductants for U(VI) or Tc(VI).2−5 Reactions involved in the oxidation of reduced Fe/S minerals are also capable of altering the behavior of redoxsensitive metal−radionuclide contaminants, such as U and Tc, potentially leading to increased solubility and mobilization of these contaminants.6,7 Previous and ongoing studies indicate that lithotrophic microorganisms are capable of oxidizing insoluble reduced Fe/S phases and minerals, such as biotite, smectite, iron monosulfide, and pyrite, with oxygen or nitrate as the electron acceptor.8−14 Understanding the capacity of native subsurface microbial communities to oxidize reduced Fe and S minerals in response to oxidant flux is thus important and crucial to predict long-term contaminant stability. The U.S. Department of Energy’s (DOE’s) Hanford 300 Area (300A) site (formerly 300 Area South Process Pond) contains significant U and nitrate groundwater contamination stemming from leakage of nuclear enrichment facilities. Seasonal fluctuations of 1−3 m in the adjacent Columbia © 2014 American Chemical Society

river serve as a source to introduce chemical oxidants to subsurface sediments,15 potentially stimulating native lithotrophic microbial communities.14 Hanford 300 Area sediments are composed of the Miocene−Pliocene-aged lacustrine Ringold formation overlain by the course-grained Pleistoceneaged Hanford formation.16,17 A discrete redox transition zone exists at 17−18 m depth in upper Ringold formation sediments (0.5−1 m below the Hanford−Ringold contact), which has been the subject of recent work focused on geochemical and microbial interactions.14,15,17,18 Fe in reduced Ringold sediment exists as predominantly phyllosilicates and (oxyhydr)oxides, with smaller quantities of magnetite, siderite, and pyrite also present.3,18 The presence of microbial Fe(III) reduction in anoxic Ringold sediments is consistent with the relatively high abundance of Geobacteraceae-related 16S rRNA genes in the vicinity of the redox transition in the upper Ringold formation15 and abundance of Fe(III) in the vicinity of the redox transition.18 Although previous work has recovered Fe(II)-oxidizing microbes from the Hanford 300 Area subsurface redox transition zone,14 the ability of native Fe(II)-bearing Received: Revised: Accepted: Published: 9197

February 26, 2014 July 8, 2014 July 11, 2014 July 11, 2014 dx.doi.org/10.1021/es5009856 | Environ. Sci. Technol. 2014, 48, 9197−9204

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Figure 1. HCl-extractable Fe(II) (0.5 M) concentrations over time in experimental reactors with reduced Ringold sediment. Transition Inoc and reduced Inoc referred to non-sterilized reactors inoculated with fresh (undried) material from within and below the transition zone in Hanford 300 Area sediments, respectively. All values represent the mean ± range of duplicate reactors. Dashed lines indicate the average Fe(II) content of a given set of reactors during the first few weeks of incubation. Stars indicate time points where DNA was recovered for microbial community analysis.

with oxygen-free N2 for 60 min, sealed with thick rubber stoppers, and crimped to seal. All additions and removals from experimental reactors were performed using syringes with 18G needles flushed with oxygen-free N2, and all solutions added to these reactors were made anoxic and sterile by bubbling with oxygen-free N2 and autoclaving. The Hanford AGW was modified from Zachara et al.20 and formulated to mimic the geochemistry of Hanford 300 Area groundwater. Constituents (per liter) were 0.121 g of NaHCO3, 0.016 g of KHCO3, 0.0614 g of MgSO4, 0.0163 g of CaSO4, 0.0543 g of Ca(NO3)2· 4H2O, 0.0956 g of CaCl2·2H2O, and 3.35 g of piperazine-N,Nbis-2-ethanesulfonic acid (PIPES buffer). Ca(NO3)2·4H2O was replaced with CaCl2·2H2O for oxidant-free slurries. Non-sterile reactors were inoculated (5%, vol/vol) with a 10% (wt/vol) slurry of fresh (undried) sediment from either within or below the redox transition and AGW. Bottles serving as sterile controls underwent pasteurization at 80 °C for 10 min on 3 non-consecutive days, with 48−72 h between pasteurization events. Nitrate-containing reactors were amended with 5 mM NaNO3, whereas the headspace of aerobic reactors was flushed with sterile air for 5 min. All reactor treatments were prepared in duplicate. Bottles were incubated sideways in the dark at room temperature on a rotary shaker set at 50 rpm. Reactor Sampling and Geochemical Analysis. Reactors were sampled every 3−7 days for the first 3 months, every 14− 56 days for a 234 day incubation period, and again at day 910. An additional experiment designed to directly document organic carbon (OC) oxidation was sampled every ca. 14 days over a ca. 85 day period. During sampling, 2 mL subsamples were removed from each reactor with a N2-flushed syringe and needle, transferred to an anaerobic chamber, and spun in a microcentrifuge. This procedure eliminated the potential for H2 contamination of the reactors from the anaerobic chamber. The aqueous phase was used to determine

minerals to support these oxidative transformations is poorly resolved. This study sought to interrogate the potential for biotic and abiotic oxidation of native reduced Fe/S phases in Ringold sediments from beneath the transition zone in response to the input of oxygen or nitrate. The work was motivated by the recent enrichment and isolation of aerobic and nitrate-reducing organisms from Hanford 300 Area groundwater, which are capable of oxidizing solid-phase Fe(II) minerals, such as biotite and reduced smectite.14 Shifts in microbial community composition accompanying oxidant exposure were monitored through community 16S rRNA gene pyrosequencing. The results suggest that heterotrophic carbon metabolism is likely to play a key role in governing Fe/S redox speciation and the stability of the redox boundary in situ.

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MATERIALS AND METHODS Sediment Collection and Preparation. Subsurface sediment was obtained from the U.S. DOE’s Hanford 300 Area Integrated Field Research Challenge (IFRC) site (http:// ifchanford.pnl.gov), borehole 399-2-25, well IC C7870,19 in March 2011. Sediment from the redox transition was recovered aseptically from ca. 17.8 m depth and reduced material from ca. 18.5 m depth of cores recovered during drilling. Material was stored under anoxic conditions at 4 °C until use. A quantity of reduced Ringold sediment was dried anoxically and ground with a mortar and pestle in an anaerobic chamber until it passed through a 0.5 mm sieve. This ensured even particle size distribution, and testing of several water/sediment mixtures showed it was possible to obtain uniform subsamples of the suspended solids via needle and syringe. Reactor Construction. Sieved reduced Ringold sediment (10 g) was added to 100 mL of artificial Hanford groundwater (AGW) in 250 mL serum bottles. The bottles were bubbled 9198

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ca. 30% of total Fe(II) loss in the sterile oxic reactors and ca. 50% of Fe(II) increase in the oxidant-free inoculated reactors (data not shown). A distinct color change from bluish green to brown was observed in the sterile oxic reactors (see Figure S1A of the Supporting Information), whereas no change in color was observed in oxidant-free, aerobic, or nitrate-amended reactors (see panels B−D in Figure S1 of the Supporting Information). pH values in all reactors remained between 6.8 and 7.1, regardless of inoculation or oxidant treatment. Sulfate concentrations in oxidant-free inoculated reactors showed total sulfate consumption by the end of the experiment. Sulfate levels remained constant in uninoculated oxidant-free reactors (Figure 2A) as well as in all nitrate-amended reactors,

sulfate, nitrite, and nitrate concentrations using ion chromatography (Dionex model 1000). The solid phase was rinsed 3 times with anoxic AGW to remove any residual nitrite and subjected to a 24 h 0.5 M HCl extraction for quantification of Fe(II) and total Fe. Fe was measured colorimetrically using ferrozine with and without the addition of hydroxylamine−HCl in duplicate.21 When both nitrate and nitrite concentrations fell to concentrations below 2 mM in a given nitrate-amended reactor, an additional ca. 8 mM nitrate was added from a sterile, anaerobic, stock solution. At these same time points, the headspaces of all aerobic reactors, including sterile controls, were reflushed with sterile air for 5 min. During the OC oxidation experiment, headspace CO2 and dissolved inorganic carbon concentrations were determined by gas chromatography as previously described.22 The initial total reduced inorganic sulfur content of the reduced Ringold sediment was measured via the single-step chromium reduction assay23 with colorimetric sulfide determination.24 The total Fe and Fe(II) contents of the reactors at the end of experiments was determined by the HF-1,10phenanthroline (HF-phenanthroline) method, which quantitatively recovers Fe from both Fe(III) oxides and Fe(III)-bearing phyllosilicates.25 The HF-phenanthroline method used was a modification of the method described by Komadel and Stucki,25 in which hydroxylamine sulfate (rather than light) was used to reduce all Fe in the extract for determination of total Fe.26 DNA Extraction and Analysis. At days 0, 34, 76, and 104, 2 mL samples were collected from all reactors for DNA extraction. These samples were immediately frozen at −80 °C until processing. DNA was extracted from thawed samples with a Mo-Bio PowerSoil DNA Isolation kit. Purified DNA was quantified using a Qubit fluorometer (Invitrogen) following the protocols of the manufacturer. High-throughput barcode sequencing was completed at the University of Wisconsin Madison on the Roche GS Junior platform in December 2011. Primers 515F and 806R27 were used to target the V4 region of the 16s rRNA gene. After analysis, reads were checked for errors and chimeras and analyzed using quantitative insights into microbial ecology (QIIME),28 with operational taxonomic units (OTUs) calculated based on 99% similarity. Taxonomy was assigned though QIIME, and BLAST associations29 were also investigated for high-abundance OTUs. After chimera and error checking, an average of 3500 reads/sample was returned (range of 919−25 737 reads/sample). Because of the variable number of reads, all taxonomic data are presented as individual hits to a given OTU, normalized to total returned reads for a given sample. OTUs with abundance of less than 0.1% of total reads were not considered.

Figure 2. Sulfate concentrations over time in experimental reactors with reduced Ringold sediment. No inoc refers to sterile, uninoculated reactors, and trans inoc and red inoc refer to non-sterilized reactors inoculated with fresh (undried) material from within and below the transition zone in Hanford 300 Area sediments, respectively. All values represent the mean ± range of duplicate reactors.

regardless of inoculation (Figure 2B). Inoculated oxic reactors showed no change in sulfate concentrations, whereas sterile oxic reactors showed a small but detectable increase from ca. 0.5 to 0.7 mM sulfate during the experiment (Figure 2C). Nitrate Consumption in Nitrate-Amended Reactors. Significant nitrate consumption was observed in inoculated nitrate-amended reactors, whereas sterile reactors showed no change in the nitrate concentration (Figure 3). The initial nitrate additions were completely consumed by day 14 in reactors inoculated with material from the transition zone as well as one duplicate inoculated with reduced material. No nitrite was detected during this initial period of nitrate consumption. An additional ca. 5 mM nitrate added to these reactors was consumed within 14 days, and additional nitrate was again added to all inoculated reactors on days 27 and 48. The rate of nitrate consumption declined after day 48, and nitrite concentrations increased to 2−5 mM in all reactors. At day 140, nitrite concentrations in three reactors fell below 2

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RESULTS Fe(II) Oxidation and Sulfate Generation. Significant oxidation of 0.5 M HCl-extractable Fe(II) was observed only in sterile, uninoculated oxic reactors, in which Fe(II) levels decreased from 10−15 to ca. 5 mmol L−1 over the 910 day incubation period (Figure 1G). Neither the inoculated oxic reactors nor any of the sterile or inoculated nitrate-amended reactors showed systematic Fe(II) loss (panels D, E, F, H, and I of Figure 1). Treatments with no added oxidant likewise showed no Fe(II) loss (panels B and C of Figure 1), and a noticeable increase (5−10 mmol L−1) in Fe(II) was noted in inoculated reactors (Figure 1A). HF-phenanthroline analysis of the total Fe(II) content of reactor solids at the end of the experiment indicated that the 0.5 M HCl extraction captured 9199

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oxidants and microbially and/or chemically driven reductive processes. This balance is important to the fate of subsurface contaminants, particularly nitrate and U, because the redox boundary could serve as a vertical barrier that mitigates widespread contaminant mobilization.18 The goal of our experiments was to assess the ability of native microbial communities to oxidize sediment-associated reduced Fe/S phases in response to oxygen and nitrate influx and, thereby, potentially influence redox balance within the transition zone. Oxidant-Free Reactors. Reactors with no oxidant addition provided a negative control with respect to potential reduced Fe/S oxidation. The uninoculated reactors showed no systematic change in Fe(II) (Figure 1A) or sulfate (Figure 2A) concentration, indicating stable anoxic conditions. The inoculated reactors showed evidence of both Fe(II) production (panels B and C of Figure 1) and sulfate consumption (Figure 2A). Some of the HCl-extractable Fe(II) production could have been due to the reaction of sulfide, produced during bacterial sulfate reduction, with oxide- and/or phyllosilicate-associated Fe(III).30 These results suggest the presence of active anaerobic heterotrophic communities in the Ringold sediment, which is consistent with previous molecular- and cultivation-based studies with sediment from within15 and above17 the redox transition. The metabolism of these organisms may help to maintain reducing conditions below the redox boundary. Nitrate-Amended Reactors. Inoculated, nitrate-amended reactors showed significant nitrate consumption (Figure 3), whereas nitrate consumption was absent in uninoculated reactors. These results are indicative of biological nitrate reduction, which agrees with the development of a large overpressure of (presumably) N2 gas in the headspace of the inoculated reactors. Despite extensive nitrate reduction, no oxidation of reduced Fe or S was apparent. These results indicate that nitrate reduction was driven by oxidation of OC rather than reduced Fe/S phases. Consistent with this suggestion, microbial communities in the nitrate-amended reactors were dominated by known heterotrophic nitratereducing taxa from the family Pseudomonadaceae31 (Figure 4). Recent studies of redox transition sediments have identified other heterotrophic nitrate-reducing taxa,15 and thus, our finding of significant nitrate reduction potential in Ringold sediments was anticipated. However, the lack of reduced Fe/S oxidation was unexpected in light of the documented presence of a chemolithotrophic Fe(II)-oxidizing, nitrate-reducing organism from Hanford 300 Area redox transition zone sediments.14 The simplest interpretation of these results is that heterotrophic metabolism retarded potential nitrate-driven reduced Fe/S oxidation in our experiments by consuming available nitrate. The lack of significant reduced Fe/S oxidation toward the end of the experiment when rates of (presumed) OC-driven nitrate/nitrite consumption declined is puzzling but may be attributed simply to a low abundance of lithotrophic microbial taxa. Oxygen-Containing Reactors. Similar to the nitrateamended sediments, inoculated reactors with an aerobic headspace showed no significant oxidation of reduced Fe or S phases. Microbial communities were dominated by Pseudomonadaceae and other heterotrophic taxa, including anaerobic (fermentative) organisms from the family Clostridiaceae (Figure 4). These results suggest that the rate of reactor shaking was too low to overcome oxygen limitation induced by aerobic heterotrophs rapidly consuming dissolved oxygen and creating conditions favorable for growth of anaerobic taxa. The

Figure 3. Nitrate and nitrite concentrations over time in nitrateamended experimental reactors with reduced Ringold sediment. No inoc refers to sterile, uninoculated reactors, and trans inoc and red inoc refer to non-sterilized reactors inoculated with fresh (undried) material from within and below the transition zone in Hanford 300 Area sediments, respectively. All values represent the mean ± range of duplicate reactors.

mM, at which time a final addition of 10 mM nitrate was made. In total, consumption of nitrate ranged from 17 to 22 mM in the four inoculated reactors. Microbial Community Composition. Significant changes to microbial communities occurred during incubation, with oxidant treatment, rather than initial inoculum source, being the strongest driver of community composition (Figure 4). Initial material from both sources had similar community composition; however, relative abundance was very different: the transition material was dominated by Pseudomonadaceae, whereas the reduced material was dominated by Burkholderiaceae. Establishment of communities was rapid following inoculation, and convergence occurred by day 34 in most reactors, with these communities remaining stable for the duration of sampling (≥100 days). Not surprisingly, communities shifted the least in reactors with no additional oxidant, which were dominated by Pseudomonadaceae (putatively Pseudomonas aeruginosa), Rhizobiaceae (Bradyrhizobium sp.), and Clostridiaceae. Oxic reactors showed the most difference as a function in inoculum source, yet both contained high abundances of Rhodocyclaceae (Dechloromonas sp.) and Pseudomonadaceae (Pseudomonas fluorescens and P. aeruginosa). Clostridiaceae was only observed in rectors inoculated with reduced material. Additional assorted populations (2 million years old) crystalline reduced Fe phases in native reduced Ringold sediments are less reactive toward oxygen compared to the new Fe(II) phases produced during microbial reduction experiments. The important point is that, as in the case of the nitrate-amended reactors, we infer that the metabolism of aerobic heterotrophic organisms retarded potential biotic and/or abiotic oxidation of reduced Fe/S phases in the inoculated reactors. A follow-up experiment verified generation of inorganic carbon coupled to OC oxidation in oxygen- and nitrate-amended sediment slurries (see Figure S3 of the Supporting Information). Implications for the Maintenance of the Subsurface Redox Boundary. The observed potential for consumption of

absence of major proliferation of Clostridiaceae in the nitrateamended reactors is consistent with this inference, where denitrifiers would be expected to outcompete fermentative and other anaerobic respiratory organisms in the presence of millimolar nitrate concentrations.32 Sterile oxic treatments were the only reactors to experience significant Fe/S oxidation, with a loss of 5−10 mmol L−1 Fe(II) and generation of ca. 0.2 mM sulfate. These changes corresponded to oxidation of ca. 50% of 0.5 M HCl-extractable Fe(II) and ca. 10% of the reduced inorganic sulfur pool. The loss of 0.5 M HCl-extractable Fe(II) in these reactors was paralleled by a decline in total 0.5 M HCl-extractable Fe (see Figure S2 of the Supporting Information). This result suggests that the majority of Fe(II) oxidized was structural Fe in phyllosilicates, which is much more soluble (in dilute HCl) in Fe(II) compared to the Fe(III) redox state.33 It is also possible that some non-phyllosilicate Fe(II) [e.g., in the form of sorbed Fe(II) or siderite] was oxidized to crystalline Fe(III) oxides, which are only poorly soluble in dilute HCl,34 as observed in 9201

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Figure 5. Diffusion−reaction simulation of oxygen and nitrate consumption coupled to OC oxidation in Ringold formation sediments, using a firstorder OC oxidation rate constant (kOC) of 7.3 ± 10−6, 7.3 ± 10−5, or 7.3 ± 10−7 year−1. Symbols show observed oxygen and nitrate concentrations in the vicinity of the redox boundary,15 and lines show simulation results at five time points over a 15 000 year simulation.

end of values known for deeply buried marine sediments.41 Rate constants 10-fold lower or higher than this value failed to reproduce the depth distributions of oxygen and nitrate (panels B and C of Figure 5); alteration of the “limiting” oxygen and nitrate concentrations by a factor of 10 had little impact on the simulated depth profiles. The profiles showed only a minor change during the simulation period (Figure 5A), which makes sense given that only ca. 10% of the OC pool was consumed (data not shown). Additional simulations (not shown) indicate that the redox barrier is likely to be stable in its current location for several tens of thousands of years. It should be noted that the rates of OC oxidation in the homogenized sediment slurries were several orders of magnitude greater than those inferred from the model, presumably because of sediment disaggregation and exposure of reactive solid-phase organic carbon to microorganisms and oxidants, as documented previously for deeply buried clay-rich Atlantic coastal plain sediments.42 The model did not include consumption of oxygen or nitrate coupled to the oxidation of reduced Fe/S phases and, thus, does not fully reproduce the geochemistry of the redox transition zone, where a significant decline in the abundance of reduced Fe and S occurs over a ca. 0.5 m depth interval.17,18 A distinct color change from light tan to greenish blue takes place across this interval,43 which is reminiscent of the difference in color between solids in the sterile, oxygen-amended reactors versus all other reactors at the end of our experiments (see Figure S4 of the Supporting Information). Assuming that no OC oxidation took place in the sterile, oxygen-amended reactors (supported by the absence of nitrate consumption in the sterile nitrate-amended reactors; see Figure 3), this similarity suggests that the oxidation of reduced Fe/S that has taken place above the redox boundary can be attributed to at least partial exhaustion of metabolizable OC within this layer over the past 10 000−15 000 years. This assertion is supported by measurements of the OC content of upper Ringold formation sediments (McKinley et al., unpublished data), which increase from values of ≤0.03 to 0.1−0.8% dry weight with depth going across the redox transition zone. Progressive depletion of OC with depth could have altered the balance between heterotrophic microbial oxidant use versus biotic and/ or abiotic consumption coupled to reduced Fe/S oxidation in a manner analogous to how sterilization of the reduced Ringold sediment rendered reduced Fe/S phases susceptible to oxidation by oxygen. Mixing of reduced Fe/S-poor materials into the upper few decimeters of Ringold formation sediment during deposition of the Hanford formation could also help to account for the presence of the oxidized layer just below the Hanford−Ringold contact. In any case, a key reality is that the layer of fine-grained reduced sediment beneath the redox transition is ca. 3 m deep,43 which means that the redox

oxidants coupled to heterotrophic microbial metabolism suggests that OC oxidation is likely to play an important role in the maintenance of the redox boundary in upper Ringold formation sediments. The OC content of the reduced Ringold sediment ranges from ca. 0.1 to 1% by weight,18 sufficient to account for the quantities of nitrate (and presumably oxygen) consumed in our experiments. A diffusion/reaction model was developed to assess whether consumption of oxygen and nitrate coupled to OC oxidation could maintain the redox boundary over a 10 000 year time frame. The model is based on the assumption that buried, Pliocene-aged fine-grained reduced Ringold formation sediments became exposed to oxygen and nitrate input subsequent to catastrophic deposition of the coarse-grained overlying Pleistocene-aged Hanford formation sediments.16 A 2 m deep reaction zone was adopted on the basis of the observed spatial scale of groundwater oxygen and nitrate concentrations in the vicinity of the redox boundary in Hanford 300 Area sediments.15 Initial concentrations of oxygen and nitrate were set equal to zero under the assumption that the exposed sediments were initially anoxic and nitrate-free, as is the case for sediments below the current redox boundary.15 The starting sediment OC content was set equal to 0.1%, and a porosity of 30% was assumed for the mud facies present in the upper Ringold.36 The model assumed that solute transport was diffusioncontrolled, consistent with the fine-grained nature of these materials compared to the much more transmissive overlying Hanford formation sediments.15 Diffusive transport was simulated via finite differences (200 node points) using the numerical method of lines.37 Diffusion coefficients for nitrate and oxygen were calculated using the DIFCOEF algorithm with a temperature of 15 °C and corrected for tortuosity by multiplying by porosity squared.37 OC oxidation was assumed to be a first-order reaction, and consumption of oxygen and nitrate coupled to OC oxidation was modeled via the modified Monod approach38−40 (see Table S1 of the Supporting Information) with oxygen as the preferred electron acceptor;31 for simplicity, the “limiting” oxygen and nitrate concentrations were both set equal to 10 μM. The rate constant for OC oxidation was adjusted to reproduce the approximate gradients in oxygen and nitrate across the current redox boundary. The results of the simulation support the idea that OC oxidation could have exerted primary control on the redox boundary over the past ca. 10 000 years (Figure 5). Testing showed that a period of ca. 5000 years was required to produce uniform concentrations of oxygen and nitrate in the absence of their consumption via OC oxidation; hence, a total simulation period of 15 000 years was used to simulate a full 10 000 years of coupled diffusion and reaction. The inferred first-order rate constant for OC oxidation (7.3 × 10−6 year−1) is near the low 9202

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(2) Jeon, B. H.; Barnett, M. O.; Burgos, W. D.; Dempsey, B. A.; Roden, E. E. Chemical reduction of U(VI) by Fe(II) at the solid− water interface using synthetic and natural iron(III) oxides. Environ. Sci. Technol. 2005, 39, 5642−5649. (3) Fredrickson, J. K.; Zachara, J. M.; Kennedy, D. W.; Kukkadapu, R. K.; McKinley, J. P.; Heald, S. M.; Liu, C.; Plymale, A. E. Reduction of TcO4− by sediment-associated biogenic Fe(II). Geochim. Cosmochim. Acta 2004, 68, 3171−3187. (4) Hyun, S. P.; Davis, J. A.; Sun, K.; Hayes, K. F. Uranium(VI) reduction by iron(II) monosulfide mackinawite. Environ. Sci. Technol. 2012, 46 (6), 3369−3376. (5) Lee, J. H.; Zachara, J. M.; Fredrickson, J. K.; Heald, S. M.; McKinley, J. P.; Plymale, A. E.; Resch, C. T.; Moore, D. A. Fe(II)- and sulfide-facilitated reduction of 99Tc(VII)O4− in microbially reduced hyporheic zone sediments. Geochim. Cosmochim. Acta 2014, 136, 247− 264. (6) Komlos, J.; Peacock, A.; Kukkadapu, R. K.; Jaffe, P. R. Long-term dynamics of uranium reduction/reoxidation under low sulfate conditions. Geochim. Cosmochim. Acta 2008, 72 (15), 3603−3615. (7) Fredrickson, J. K.; Zachara, J. M.; Plymale, A. E.; Heald, S. M.; McKinley, J. P.; Kennedy, D. W.; Liu, C. X.; Nachimuthu, P. Oxidative dissolution potential of biogenic and ablogenic TcO2 in subsurface sediments. Geochim. Cosmochim. Acta 2009, 73 (8), 2299−2313. (8) Shelobolina, E. S.; Vanpraagh, C. G.; Lovley, D. R. Use of ferric and ferrous iron containing minerals for respiration by Desulfitobacterium frappieri. Geomicrobiol. J. 2003, 20 (2), 143−156. (9) Haaijer, S. C. M.; Lamers, L. P. M.; Smolders, A. J. P.; Jetten, M. S. M.; den Camp, H. J. M. O. Iron sulfide and pyrite as potential electron donors for microbial nitrate reduction in freshwater wetlands. Geomicrobiol. J. 2007, 24 (5), 391−401. (10) Torrentó, C.; Cama, J.; Urmeneta, J.; Otero, N.; Soler, A. Denitrification of groundwater with pyrite and Thiobacillus denitrificans. Chem. Geol. 2010, 278 (1−2), 80−91. (11) Bosch, J.; Lee, K. Y.; Jordan, G.; Kim, K. W.; Meckenstock, R. U. Anaerobic, nitrate-dependent oxidation of pyrite nanoparticles by Thiobacillus denitrificans. Environ. Sci. Technol. 2012, 46 (4), 2095− 2101. (12) Shelobolina, E.; Konishi, H.; Xu, H.; Benzine, J.; Xiong, M. Y.; Wu, T.; Blothe, M.; Roden, E. Isolation of phyllosilicate−iron redox cycling microorganisms from an illite−smectite rich hydromorphic soil. Front. Microbiol. 2012, 3, 134. (13) Shelobolina, E.; Xu, H.; Konishi, H.; Kukkadapu, R.; Wu, T.; Blothe, M.; Roden, E. Microbial lithotrophic oxidation of structural Fe(II) in biotite. Appl. Environ. Microbiol. 2012, 78 (16), 5746−5752. (14) Benzine, J.; Shelobolina, E. S.; Xiong, M. Y.; Kennedy, D. W.; McKinley, J. P.; Lin, X.; Roden, E. E. Fe-phyllosilicate redox cycling organisms from a redox transition zone in Hanford 300 Area sediments. Front. Microbiol. 2013, 4, 388. (15) Lin, X. J.; Kennedy, D.; Peacock, A.; McKinley, J.; Resch, C. T.; Fredrickson, J.; Konopka, A. Distribution of microbial biomass and potential for anaerobic respiration in Hanford Site 300 Area subsurface sediment. Appl. Environ. Microbiol. 2012, 78 (3), 759−767. (16) Zachara, J. M.; Serne, J.; Freshley, M.; Mann, F.; Anderson, F.; Wood, M.; Jones, T.; Myers, D. Geochemical processes controlling migration of tank wastes in Hanford’s vadose zone. Vadose Zone J. 2007, 6 (4), 985. (17) Lee, J. H.; Fredrickson, J. K.; Kukkadapu, R. K.; Boyanov, M. I.; Kemner, K. M.; Lin, X. J.; Kennedy, D. W.; Bjornstad, B. N.; Konopka, A. E.; Moore, D. A.; Resch, C. T.; Phillips, J. L. Microbial reductive rransformation of phyllosilicate Fe(III) and U(VI) in fluvial subsurface sediments. Environ. Sci. Technol. 2012, 46 (7), 3721−3730. (18) Peretyazhko, T. S.; Zachara, J. M.; Kukkadapu, R. K.; Heald, S. M.; Kutnyakov, I. V.; Resch, C. T.; Arey, B. W.; Wang, C. M.; Kovarik, L.; Phillips, J. L.; Moore, D. A. Pertechnetate (TcO4−) reduction by reactive ferrous iron forms in naturally anoxic, redox transition zone sediments from the Hanford Site, USA. Geochim. Cosmochim. Acta 2012, 92, 48−66. (19) Bjornstad, B.; Lanigan, D.; Horner, J.; Thorne, P.; Vermeul, V. Borehole Completion and Conceptual Hydrogeologic Model for the IFRC

boundary is likely to persist over geological time scales. Simulations made with a 4 m deep reaction zone suggest that the redox boundary would be expected to migrate downward ca. 1 m over a 200 000 year time period. In summary, this work documented the chemical and microbiological responses of reduced Ringold formation sediments from the Hanford 300 Area site to the influx of nitrate or oxygen. In contrast to the initial hypothesis that oxidation of reduced Fe/S phases by native microbial communities could exert a dominant control on sediment redox status, Fe(II) oxidation (and accompanying modest sulfate generation) was only observed in aerobic, sterile controls. All other inoculated and sterile reactors amended with oxygen or nitrate showed no Fe(II) loss or sulfate generation. Non-sterile sediments consumed significant amounts of nitrate (and putatively oxygen) and were dominated by heterotrophic microbes consuming residual sediment OC. Heterotrophic metabolism consumed available oxidants, thus preventing biotic and/or abiotic reduced Fe/S oxidation. These findings are consistent with previous observations that Ringold formation sediments contain active heterotrophic microbial communities15 and reveal that such communities are poised to respond to oxidant influx and, thereby, act as a barrier to vertical U, nitrate, and other redoxsensitive contaminant migration over millennial time scales.

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ASSOCIATED CONTENT

S Supporting Information *

Modified Monod simulation of oxygen and nitrate consumption (Table S1), photos of sediment reactors at the end of the incubation experiments (Figure S1), total 0.5 M HClextractable Fe concentrations over time in oxygen-containing reactors (Figure S2), accumulation of total dissolved inorganic carbon in aerobic and nitrate-amended reactors compared to no oxidant controls (Figure S3), and a photo of upper Ringold formation core section illustrating the color change that takes place across the redox transition zone, in comparison to the color of solids in the sediment reactors (Figure S4). This material is available free of charge via the Internet at http:// pubs.acs.org.

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AUTHOR INFORMATION

Corresponding Author

*Telephone: +1-608-831-3680. E-mail: [email protected]. edu. Notes

The authors declare no competing financial interest.

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ACKNOWLEDGMENTS The authors thank B. J. Converse for assistance with the 16S rRNA gene sequencing and analysis and J. K. Fredrickson [Pacific Northwest National Laboratory (PNNL)] and C. Liu (PNNL) for review of the paper. This work was supported by the U.S. DOE, Office of Biological and Environmental Research, Subsurface Biogeochemical Research (SBR) Program through the SBR Scientific Focus Area at PNNL.

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REFERENCES

(1) Borch, T.; Kretzschmar, R.; Kappler, A.; Van Cappellen, P.; Ginder-Vogel, M.; Voegelin, A.; Campbell, K. Biogeochemical redox processes and their impact on contaminant dynamics. Environ. Sci. Technol. 2010, 44 (1), 15−23. 9203

dx.doi.org/10.1021/es5009856 | Environ. Sci. Technol. 2014, 48, 9197−9204

Environmental Science & Technology

Article

Well Field, 300 Area, Hanford Site; Pacific Northwest National Laboratory (PNNL): Richland, WA, 2009. (20) Zachara, J. M. Multi-scale Mass Transfer Processes Controlling Natural Attenuation and Engineered Remediation: An IFRC Focused on Hanford’s 300 Area Uranium Plume (January 2008 to January 2009); Pacific Northwest National Laboratory (PNNL): Richland, WA, 2009. (21) Stookey, L. FerrozineA new spectrophotometric reagent for iron. Anal. Chem. 1970, 42 (7), 779−781. (22) Roden, E. E.; Wetzel, R. G. Organic carbon oxidation and suppression of methane production by microbial Fe(III) oxide reduction in vegetated and unvegetated freshwater wetland sediments. Limnol. Oceanogr. 1996, 41, 1733−1748. (23) Fossing, H.; Jorgensen, B. B. Measurement of bacterial sulfate reduction in sediments: Evaluation of a single-step chromium reduction method. Biogeochemistry 1989, 8, 205−222. (24) Cline, J. D. Spectrophotometric determination of hydrogen sulfide in natural waters. Limnol. Oceanogr. 1969, 14, 454−458. (25) Komadel, P.; Stucki, J. W. Quantitative assay of minerals for Fe2+ and Fe3+ using 1,10-phenanthroline: III. A rapid photochemical method. Clays Clay Miner. 1988, 36 (4), 379−381. (26) Amonette, J. E.; Templeton, J. C. Improvements to the quantitative assay of nonrefractory minerals for Fe(II) and total Fe using 1,10-phenanthroline. Clays Clay Miner. 1998, 46 (1), 51−62. (27) Caporaso, J. G.; Lauber, C. L.; Walters, W. A.; Berg-Lyons, D.; Lozupone, C. A.; Turnbaugh, P. J.; Fierer, N.; Knight, R. Global patterns of 16S rRNA diversity at a depth of millions of sequences per sample. Proc. Natl. Acad. Sci. U. S. A. 2011, 108, 4516−4522. (28) Caporaso, J. G.; Kuczynski, J.; Stombaugh, J.; Bittinger, K.; Bushman, F. D.; Costello, E. K.; Fierer, N.; Pena, A. G.; Goodrich, J. K.; Gordon, J. I.; Huttley, G. A.; Kelley, S. T.; Knights, D.; Koenig, J. E.; Ley, R. E.; Lozupone, C. A.; McDonald, D.; Muegge, B. D.; Pirrung, M.; Reeder, J.; Sevinsky, J. R.; Tumbaugh, P. J.; Walters, W. A.; Widmann, J.; Yatsunenko, T.; Zaneveld, J.; Knight, R. QIIME allows analysis of high-throughput community sequencing data. Nat. Methods 2010, 7 (5), 335−336. (29) Altschul, S. F.; Madden, T. L.; Schäffer, A. A.; Zhang, J.; Zhang, Z.; Miller, W.; Lipman, D. J. Gapped BLAST and PSI-BLAST: A new generation of protein database search programs. Nucleic Acids Res. 1997, 25 (17), 3389. (30) Poulton, S. W. Sulfide oxidation and iron dissolution kinetics during the reaction of dissolved sulfide with ferrihydrite. Chem. Geol. 2003, 202 (1−2), 79−94. (31) Tiedje, J. M. Ecology of denitrification and dissimilatory nitrate reduction to ammonium. In Biology of Anaerobic Microorganisms; Zehnder, A. J. B., Ed.; John Wiley and Sons: New York, 1988. (32) Sorensen, J. Reduction of ferric iron in anaerobic, marine sediment and interaction with reduction of nitrate and sulfate. Appl. Environ. Microbiol. 1982, 43, 319−324. (33) Kostka, J. E.; Stucki, J. W.; Nealson, K. H.; Wu, J. Reduction of structural Fe(III) in smectite by a pure culture of Shewanella putrefaciens strain MR-1. Clays Clay Miner. 1996, 44 (4), 522−529. (34) Roden, E.; Zachara, J. Microbial reduction of crystalline iron(III) oxides: Influence of oxide surface area and potential for cell growth. Environ. Sci. Technol. 1996, 30 (5), 1618−1628. (35) Weber, K. A.; Churchill, P. F.; Urrutia, M. M.; Kukkadapu, R. K.; Roden, E. E. Anaerobic redox cycling of iron by wetland sediment microorganisms. Environ. Microbiol. 2006, 8, 100−113. (36) Lindsey, K. A.; Gaylord, D. R. Lithofacies and sedimentology of the Miocene−Pliocene Ringold formation, Hanford site, south-central Washington. Northwest Sci. 1990, 64 (3), 165−180. (37) Boudreau, B. P. Diagenetic Models and Their Implementation; Springer: New York, 1997. (38) Boudreau, B. P.; Westrich, J. T. The dependence of bacterial sulfate reduction on sulfate concertation in marine sediments. Geochim. Cosmochim. Acta 1984, 48, 2503−2516. (39) VanCappellen, P.; Wang, Y. Cycling of iron and manganese in surface sediments: A general theory for the coupled transport and reaction of carbon, oxygen, nitrogen, sulfur, iron, and manganese. Am. J. Sci. 1996, 296, 197−243.

(40) Hunter, K. S.; Wang, Y.; VanCappellen, P. Kinetic modeling of microbially-driven redox chemistry of subsurface environments: Coupling transport, microbial metabolism and geochemistry. J. Hydrol. 1998, 209, 53−80. (41) Middelburg, J. J. A simple rate model for organic matter decomposition in marine sediments. Geochim. Cosmochim. Acta 1989, 53, 1577−1581. (42) Chapelle, F. H.; Lovley, D. R. Rates of microbial metabolism in deep coastal plain aquifers. Appl. Environ. Microbiol. 1990, 56, 6. (43) Williams, M.; Rockhold, M.; Thorne, P.; Chen, Y. ThreeDimensional Groundwater Models of the 300 Area at the Hanford Site, Washington State; Pacific Northwest National Laboratory (PNNL): Richland, WA, 2008; PNNL-17708.

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