Dissimilatory Bacterial Reduction Research Articles

Removal of antimonate from wastewater by dissimilatory bacterial reduction: Role of the coexisting sulfate

The priority pollutant antimony (Sb) exists primarily as Sb(V) and Sb(III) in natural waters, and Sb(III) is generally with greater mobility and toxicity than Sb(V). The bio-reduction of Sb(V) would not become a meaningful Sb-removal process unless the accumulation of produced dissolved Sb(III) could be controlled. Here, we examined the dissimilatory antimonate bio-reduction with or without the coexistence of sulfate using Sb-acclimated biomass. Results demonstrated that 0.8mM Sb(V) was almost completely bio-reduced within 20h along with 48.6% Sb(III) recovery. Kinetic parameters qmax and Ks calculated were 0.54mg-Sb mg-DW−1h−1 and 41.96mgL−1, respectively. When the concentrations of coexisting sulfate were 0.8mM, 1.6mM, and 4mM, the reduction of 0.8mM Sb(V) was accomplished within 17, 9, and 5h, respectively, along with no final Sb(III) recovery. Also, the bio-reduction of sulfate occurred synchronously. The precipitated Sb2O3 and Sb2S3 were characterized by scanning electron microscopy coupled with energy dispersive spectrometer, X-ray diffraction, and X-ray photoelectron spectroscopy. Compared with bacterial compositions of the seed sludge obtained from anaerobic digestion tank in a wastewater treatment plant, new genera of Pseudomonas and Geobacter emerged with large proportions in both Sb-fed and Sb-sulfate-fed sludge, and a small portion of sulfate reduction bacteria emerged only in Sb-sulfate-fed culture.

Removal of aromatic hydrocarbons from aquifers by oxidation coupled with dissimilatory bacterial reduction of iron

Dissimilatory iron-reducing bacteria derive energy needed for growth and function by oxidation of organic compounds or molecular hydrogen coupled with reduction of iron(III) to iron(II) in subsurface environments. In this work, we used the indicated bacteria in an anaerobic environment for oxidation of benzene, toluene, ethylbenzene, and xylenes (a mixture of o-, m-, and p-xylenes) (BETX), which are aquifer contaminants. We show that oxidation of aromatic hydrocarbons coupled with bacterial reduction of iron follows first-order kinetics. The rate constant is maximum for an aromatic hydrocarbon:Fe ratio equal to 1. The oxidation rate constant increases in the series benzene < toluene < ethylbenzene < xylenes. The aromatic hydrocarbon degradation efficiency increases as the concentration of the substrate (citric acid) increases.

Surface structure effects on direct reduction of iron oxides by Shewanella oneidensis

The atomic and electronic structure of mineral surfaces affects many environmentally important processes such as adsorption phenomena. They are however rarely considered relevant to dissimilatory bacterial reduction of iron and manganese minerals. In this regard, surface area and thermodynamics are more commonly considered. Here we take a first step towards understanding the nature of the influence of mineral surface structure upon the rate of electron transfer from Shewanella oneidensis strain MR-1 outer membrane proteins to the mineral surface and the subsequent effect upon cell “activity.” Cell accumulation has been used as a proxy for cell activity at three iron oxide single crystal faces; hematite (001), magnetite (111) and magnetite (100). Clear differences in cell accumulation at, and release from the surfaces are observed, with significantly more cells accumulating at hematite (001) compared to either magnetite face whilst relatively more cells are released into the overlying aqueous phase from the two magnetite faces than hematite. Modeling of the electron transfer process to the different mineral surfaces from a decaheme (protoporphyrin rings containing a central hexacoordinate iron atom), outer membrane-bound cytochrome of S. oneidensis has been accomplished by employing both Marcus and ab initio density functional theories. The resultant model of electron transfer to the three oxide faces predicts that over the entire range of expected electron transfer distances the highest electron transfer rates occur at the hematite (001) surface, mirroring the observed cell accumulation data. Electron transfer rates to either of the two magnetite surfaces are slower, with magnetite (111) slower than hematite (001) by approximately two orders of magnitude. A lack of knowledge regarding the structural details of the heme-mineral interface, especially in regards to atomic distances and relative orientations of hemes and surface iron atoms and the conformation of the protein envelope, precludes a more thorough analysis. However, the results of the modeling concur with the empirical observation that mineral surface structure has a clear influence on mineral surface-associated cell activity. Thus surface structure effects must be accounted for in future studies of cell-mineral interactions.

Structural and compositional evolution of Cr/Fe solids after indirect chromate reduction by dissimilatory iron-reducing bacteria

The mobility and toxicity of Cr within surface and subsurface environments is diminished by the reduction of Cr(VI) to Cr(III). The reduction of hexavalent chromium can proceed via chemical or biological means. Coupled processes may also occur including reduction via the production of microbial metabolites, including aqueous Fe(II). The ultimate pathway of Cr(VI) reduction will dictate the reaction products and hence the solubility of Cr(III). Here, we investigate the fate of Cr following a coupled biotic–abiotic reduction pathway of chromate under iron-reducing conditions. Dissimilatory bacterial reduction of two-line ferrihydrite indirectly stimulates reduction of Cr(VI) by producing aqueous Fe(II). The product of this reaction is a mixed Fe(III)-Cr(III) hydroxide of the general formula Fe 1−xCr x(OH) 3 · nH 2O, having an α/β-FeOOH local order. As the reaction proceeds, Fe within the system is cycled (i.e., Fe(III) within the hydroxide reaction product is further reduced by dissimilatory iron-reducing bacteria to Fe(II) and available for continued Cr reduction) and the hydroxide products become enriched in Cr relative to Fe, ultimately approaching a pure Cr(OH) 3 · nH 2O phase. This Cr purification process appreciably increases the solubility of the hydroxide phases, although even the pure-phase chromium hydroxide is relatively insoluble.

Estimating historical in-lake alkalinity generation from sulfate reduction and its relationship to lake chemistry as inferred from algal microfossils

Sediment cores were used to estimate in-lake alkalinity generation resulting from sulfate reduction relative to inferred changes in lakewater pH and trophic status over the last century in three Connecticut lakes. Despite being situated in geological settings with crystalline bedrock and thin, poorly buffered soils, and being impacted with high rates of acidic precipitation, none of the study lakes have declined in inferred pH based on scaled chrysophyte and diatom remains. In fact, the pH of one of the lakes, Coventry Lake, has significantly increased over the last century. Over the last 44 to 69 years the amount of sulfur stored in the sediments from each lake increased from ~two to three times resulting in mean rates of alkalinity generation ranging from 78 to 145 meq m−2 yr−1, significantly higher than the 45 to 48 meq m−2 yr−1 of hydrogen ions falling directly on the lake surfaces. In-lake alkalinity generation resulting from sulfate reduction has been sufficient to neutralize all of the acid falling directly onto the lake surfaces, as well as between 9% and 25% of the acid deposited onto the surrounding watersheds. Despite the increased importance of in-lake alkalinity generation, our findings support the hypothesis that significant amounts of alkalinity are also being generated in the catchments of the study lakes. The bulk of the increases in stored sulfur in all three lakes were as Fe sulfides and not in the form of organic sulfur, suggesting that the increases were the result of dissimilatory bacterial reduction of sulfate. As a result of the large increases in storage of Fe sulfides the ratio of total iron to chromium reducible sulfur (Fe:CRS) has declined in all cores over time. Despite the overall decline in Fe:CRS in recent sediments, values are still largely above 3 in more recent sediments of two of the lakes. However, values of Fe:CRS have dropped below 1 in surface sediments of Uncas Lake, suggesting that in-lake loading of phosphorus may be responsible for a recent shift in the algal flora towards a slightly more eutrophic condition.

Dissimilatory bacterial reduction of Al-substituted goethite in subsurface sediments

The microbiologic reduction of a 0.2 to 2.0 μm size fraction of an Atlantic coastal plain sediment (Eatontown) was investigated using a dissimilatory Fe(III)-reducing bacterium (Shewanella putrefaciens, strain CN32) to evaluate mineralogic controls on the rate and extent of Fe(III) reduction and the resulting distribution of biogenic Fe(II). Mössbauer spectroscopy and X-ray diffraction (XRD) were used to show that the sedimentary Fe(III) oxide was Al-substituted goethite (13–17% Al) that existed as 1- to 5-μm aggregates of indistinct morphology. Bioreduction experiments were performed in two buffers [HCO3−; 1,4-piperazinediethansulfonic acid (PIPES)] both without and with 2,6-anthraquinone disulfonate (AQDS) as an electron shuttle. The production of biogenic Fe(II) and the distribution of Al (aqueous and sorbed) were followed over time, as was the formation of Fe(II) biominerals and physical/chemical changes to the goethite.The extent of reduction was comparable in both buffers. The reducibility (rate and extent) was enhanced by AQDS; 9% of dithionite-citrate-bicarbonate (DCB) extractable Fe(III) was reduced without AQDS whereas 15% was reduced in the presence of AQDS. XRD and Mössbauer spectroscopy were used to monitor the disposition of biogenic Fe(II) and changes to the Al-goethite. Fe(II) biomineralization was not evident by XRD. Biomineralization was observed by Mössbauer when sorbed Fe(II) concentrations exceeded a threshold value. The biomineralization products displayed Mössbauer spectra consistent with siderite FeCO3 (HCO3− buffer only) and green rust [Fe(6-x)IIFexIII(OH)12]x+[(A2−)x/2.yH2O]x−. Adsorption of biogenic Fe(II) to accessory mineral phases (e.g., kaolinite) and bacterial surfaces appeared to limit biomineralization. Al evolved during reduction was sorbed, and extractable Al increased with reduction. XRD analysis indicated that neither crystallite size or the Al content of the goethite was affected by bacterial reduction, i.e., Al release was congruent with Fe(II).

Microbial reduction of Fe(III) and sorption/precipitation of Fe(II) on Shewanella putrefaciens strain CN32.

The influence of Fe(II) on the dissimilatory bacterial reduction of an Fe(III) aqueous complex (Fe(III)-citrate(aq)) was investigated using Shewanella putrefaciens strain CN32. The sorption of Fe(II) on CN32 followed a Langmuir isotherm. Least-squares fitting gave a maximum sorption capacity of Qmax = 4.19 x 10(-3) mol/10(12) cells (1.19 mmol/m2 of cell surface area) and an affinity coefficient of log K = 3.29. The growth yield of CN32 with respect to Fe(III)aq reduction showed a linear trend with an average value of 5.24 (+/-0.12) x 10(9) cells/mmol of Fe(III). The reduction of Fe(III)aq by CN32 was described by Monod kinetics with respect to the electron acceptor concentration, Fe(III)aq, with a half-saturation constant (Ks) of 29 (+/-3) mM and maximum growth rate (micromax) of 0.32 (+/-0.02) h(-1). However, the pretreatment of CN32 with Fe(II)aq significantly inhibited the reduction of Fe(III)aq, resulting in a lag phase of about 3-30 h depending on initial cell concentrations. Lower initial cell concentration led to longer lag phase duration, and higher cell concentration led to a shorter one. Transmission electron microscopy and energy dispersive spectroscopy revealed that many cells carried surface precipitates of Fe mineral phases (valence unspecified) during the lag phase. These precipitates disappeared after the cells recovered from the lag phase. The cell inhibition and recovery mechanisms from Fe(II)-induced mineral precipitation were not identified by this study, but several alternatives were discussed. A modified Monod model incorporating a lag phase, Fe(II) adsorption, and aqueous complexation reactions was able to describe the experimental results of microbial Fe(III)aq reduction and cell growth when cells were pretreated with Fe(II)aq.

Measurement of in situ rates of selenate removal by dissimilatory bacterial reduction in sediments

vertical distribution in sediments relative to other forms • A radioisotope method for measurement of bacterial of anaerobic bacterial respiration. An understanding of respiratory reduction of selenate to elemental selenium in how the process of dissimilatory selenate reduction oper­ aquatic sediments was devised. Sediments were labeled ates in these respects should ultimately aid the design of with [75Se]selenate, incubated, and washed, and 75SeO(S) treatment schemes and the rehabilitation of selenium­ was determined as counts remaining in the sediment. Core contaminated regions. In order to answer these questions profiles of selenate reduction, sulfate reduction, and den­ and to achieve these long-term goals, it was f1I'St necessary itrification were made simultaneously in the sediments of to devise a method for measuring in situ selenate reduction. an agricultural wastewater evaporation pond. Most of the In situ measures of bacterial activities have been a focus in situ selenate reduction (850/0) and all the denitrification of microbial ecology for the past 25 years. A variety of activities were confined to the upper 4-8 em of the profile, whereas sulfate reduction was greatest below 8 em (890/0 techniques employing radioisotopes, gas chromatography, of total). The integrated areal rate of selenate reduction and/or enzyme inhibitors or analogues have been exploited was 301 J,tmol m-2 day, which results in a turnover of to assess environmental rates of methanogenesis (8), den­ water column selenate in 82.4 days. itrification (9, 10), nitrogen fixation (11), and sulfate re­ duction (12). This latter process is analogous to the methods we report herein to measure selenate reduction. Introduction However, in the case of selenate reduction, our task was Oxyanions of selenium have been identified as toxic simplified because 75Se is a 'Y-emitting radioisotope (t l / 2 constituents in drainage waters from irrigated, seleniferous =120 days), which circumvents the quenching/efficiency agricultural soils (1). This environmental problem is ap­ problem associated with liquid scintillation counting of parently quite common in many regions of the western weak ~-emitters like 35S. Thus, there was no need to United States (2). Therefore, methods that address the volatilize and trap the product 75SeO(S) as there is for question of removal of selenium oxyanions from soils [35S]sulfide, but merely to wash away the added [75Se]­ and/or drainage waters by various means have received selenate. Our results with core profiles from a selenium­ considerable attention because they offer the prospect of rich agricultural wastewater evaporation pond represent restoring water or soil quality while allowing for the con­ quasi in situ assays because they were incubated in the tinuation of irrigated farming. With regard to possible laboratory under conditions approximating the field. microbiological methods, volatilization of selenium by However, our results indicate that selenate reduction oc­ formation of alkylated gases (e.g., dimethyl selenide) has curs in the surficial layers of these sediments, in proximity been suggested for treatment of soils (3) or impacted marsh to the region of active denitrification, yet spatially sepa­ waters, such as those of the Kesterson Wildlife Refuge (4, rated from sulfate reduction occurring deeper in the core. 5). Areal rates of selenate removal obtained from this core J Recently, we reported on a novel process by which indicate its rapid removal by reduction to SeO(s), and a anaerobic bacteria respire selenate, which in turn bio­ quick turnover of water column selenate. When combined chemically reduces this oxyanion to selenite and ultimately with previous observations with regard to inhibitors and to elemental selenium (6). This dissimilatory reduction stimulators of selenium reduction, the conceptual design was demonstrated to occur in sediments, was independent of an efficient anaerobic selenate digester can be dis­ of sulfate, and was inhibited by nitrate and chromate. cerned. Bacterial cultures capable of selenate respiratory growth have been isolated (6, 7). Although we were able to show Experimental Section that this process holds the potential to rapidly remove and Sites and Sampling. Surficial (upper ro.J20 em) sedi­ sequester considerable quantities (millimolar) of added ments were collected from Hunter Drain, an agricultural selenate from sediment slurries, the questions remained drain located in Stillwater, NV, where selenium contam­ as to how rapidly it proceeds in nature, and what was its ination is documented (2), as well as from the littoral zone of Big Soda Lake, an alkaline/saline environment (13)