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Identification of a Ruminococcaceae Species as the Methyl tert-Butyl Ether (MTBE) Degrading Bacterium in a Methanogenic Consortium Tong Liu,† Hyeri Ahn,† Weimin Sun,† Lora R. McGuinness,‡ Lee J. Kerkhof,‡ and Max M. Hag̈ gblom*,† †

Department of Biochemistry and Microbiology, Rutgers University, New Brunswick, New Jersey 08901, United States Department of Marine and Coastal Science, Rutgers University, New Brunswick, New Jersey 08901, United States

‡

S Supporting Information *

ABSTRACT: The widespread use of methyl tert-butyl ether (MTBE) has caused major contamination of groundwater sources and is a concern due to its taste and odor problems, as well as its toxicity. MTBE can be degraded anaerobically which makes bioremediation of contaminated aquifers a potential solution. Nevertheless, the organisms and mechanisms that are responsible for anaerobic MTBE degradation are still unknown. The aim of our research was to identify the organisms actively degrading MTBE. For this purpose we characterized an anaerobic methanogenic culture enriched with MTBE as the sole carbon source from the New Jersey Arthur Kill intertidal strait sediment. The cultures were analyzed using stable isotope probing (SIP) combined with terminal restriction fragment length polymorphism (T-RFLP), high-throughput sequencing and clone library analysis of bacterial 16S rRNA genes. The sequence data indicated that phylotypes belonging to the Ruminococcaceae in the Firmicutes were predominant in the methanogenic cultures. SIP experiments also showed sequential incorporation of the 13C labeled MTBE by the bacterial community with a bacterium most closely related to Saccharofermentans acetigenes identified as the bacterium active in O-demethylation of MTBE. Identification of the microorganisms responsible for the activity will help us better understand anaerobic MTBE degradation processes in the field and determine biomarkers for monitoring natural attenuation.

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INTRODUCTION

compounds in groundwater aquifers and domestic/public wells.14,15 Most MTBE contaminated sites are subsurface with insignificant amounts of oxygen and hence, the fate of MTBE in the environment is mainly dependent upon anaerobic processes.16,17 Monitored natural attenuation is the most financially realistic treatment for most contaminated groundwater cases.18,19 Among different natural processes, biodegradation is the most effective method to reduce the mass of contaminants in a sustainable way.11,20,21 Anaerobic biodegradation of MTBE has been observed under a variety of different redox states, including methanogenic, sulfidogenic, Fereducing, and denitrifying conditions.22−30 Identification of the microorganisms that are capable and responsible for degrading MTBE has yet to be conclusively determined. Stable isotope probing (SIP) offers a powerful way to link specific environmental functions with the responsible organisms and to identify the key microorganisms degrading specific contaminants.31−34 DNA-SIP is based on the incorporation of a 13 C-labeled substrate into cellular nucleic acids, separation of 13 C- from 12C-DNA by density gradient centrifugation, and

Methyl tert-butyl ether (MTBE) is one of a group of oxygenates initially introduced as a fuel additive to replace tetra-ethyl lead in the late 1970s.1 MTBE enhances the octane level of gasoline and therefore improves the combustion efficiency and reduces hazardous tailpipe emissions (carbon monoxide) to the atmosphere.2 In 1990, the United States Clean Air Act amendments set the requirement for use of fuel oxygenates, and MTBE was chosen over other possible oxygenates due to its blending characteristics and reasonable price.3 As a result, up to 15% MTBE by volume was amended to gasoline, with more than 10 billion liters of MTBE produced annually between 1995 and 2002 to fulfill this requirement.2,4−6 The wide use of MTBE has caused extensive groundwater contamination.7 Drinking water that is contaminated by MTBE has unacceptable taste and odor. MTBE is also considered a potential human carcinogen and can cause reproductive mutations in aquatic wildlife.8,9 In addition to the taste, odor, and health concerns, an additional issue of MTBE is its persistence and mobility in groundwater.10−13 MTBE can be introduced into the environment from leaking underground storage tanks and pipelines, industrial plants, refueling facilities and accidental spills during transport.5,11 After several decades of heavy use, MTBE was widely spread in the United States and became one of the most frequently detected volatile organic © 2016 American Chemical Society

Received: Revised: Accepted: Published: 1455

September 28, 2015 December 12, 2015 January 4, 2016 January 4, 2016 DOI: 10.1021/acs.est.5b04731 Environ. Sci. Technol. 2016, 50, 1455−1464

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Environmental Science & Technology

methyl group to acetate, which eventually can serve as the carbon source for other members of the microbial community. To test whether acetate or biomass constituents would be used by different members of the anaerobic cultures, we set up two additional sets of SIP experiments. In the first set, 150 μM 12C or 13C-labeled acetate (both carbons labeled, Sigma, St. Louis, MO), were added to 12C MTBE amended cultures after 50% degradation had occurred. The bacterial species that can utilize acetate in the MTBE-degrading communities are likely not the primary MTBE degraders, but utilize the product of acetogenic metabolism. The community structures of acetate fed cultures were analyzed after incubation for 4 days and 1 month at 28 °C. In the second experiment, the “cross-feeding” of secondary members of the community through biomass decay was tested by providing 12C or 13C-labeled biomass (1000-fold dilution of 10X BioExpress 1000 Cell Growth Media, Cambridge Isotope Laboratories, Tewksbury, MA) to MTBE amended cultures. The community structures of biomass-fed cultures were analyzed after 14 days of incubation at 28 °C. DNA Extraction and SIP Gradients. For assessing those organisms obtaining their cellular carbon from 13C-MTBE, all SIP experiments utilized haloarchaeal 13C carrier DNA methods to enhance the separation between the active and the resident community. 42 To each SIP sample, 100 ng of 12 C Halobacterium salinarium DNA and 100 ng of 13C H. salinarium DNA was added to enhance the extraction and separation efficiency. To obtain the carrier DNA, the archaeon H. salinarium, was grown with 13C-labeled 10X BioExpress 1000 Cell Growth Media as described previously.42 13C-carrier DNA has been shown to reduce the amount of 13C-incorporation and time necessary for signal detection in a SIP experiment42 and is particularly useful for samples with limited total DNA availability. After carrier addition, genomic DNA was extracted by a modified phenol-chloroform extraction procedure to yield high-quality, minimally sheared DNA.43,44 The extracted DNA (approximately 300 ng including the MTBE community and added H. salinarium DNA) was then dissolved in a 500 μL CsCl density gradient (∼1 g mL−1) containing ethidium bromide. 13C-labeled DNA was separated from 12C DNA by 36 h of centrifugation at 225 000g in a Beckman Optima ultracentrifuge (Palo Alto, CA) using a TLA 120 rotor.42 This carrier-SIP approach yields a light 12C-DNA band, representing the resident bacteria and a heavy 13C-DNA band representing newly synthesized DNA of bacteria that incorporated either 13C-MTBE, 13C-acetate or 13C-BioExpress media. Next, 40 μL of 12C light DNA and 20 μL of 13C heavy bands were collected by pipet from the CsCl gradient under UV light visualization and samples were dialyzed (in 10 mM Tris) for 45 min.42,45 To unambiguously demonstrate 13CDNA synthesis, all 13C-carrier (heavy) bands from 12C-MTBE incubation controls needed to yield little or no amplification/ terminal-restriction fragment length polymorphism (TRFLP) signal in comparison 13C-MTBE incubations. Bacterial 16S rRNA Gene Amplification and T-RFLP Analysis. The bacterial 16S rRNA genes of genomic DNA were amplified with 20 pmol universal bacterial primers: 5′-end 6-FAM-labeled 27 forward (5′-AGAGTTTGATCCTGGCTCAG-3′) and 1100 reverse (5′-AGGGTTGCGCTCGTTG-3′) as follows: 94 °C for 5 min, followed by 94 °C for 30 s, 57 °C for 30 s, and 25 cycles with a final extension at 72 °C for 10 min (Applied Biosystems Instrument, Foster City, CA). Then, 2 μL of the first round PCR product, used as the template, was reamplified with identical PCR parameters for 30 cycles.46

identification of active populations using gene sequencing approaches.35 In this study, SIP methods were used to characterize the active microbes in methanogenic MTBE degrading enrichment cultures established with sediment from the Arthur Kill (AK) intertidal straight between New Jersey and the New York harbor.25 These methanogenic enrichment cultures were maintained through successive transfers (typically 1/5 to 1/10 dilution) for more than a decade with MTBE as the sole carbon source and reached a ∼10−10 dilution of the original inoculum. Previous studies showed that the microbial communities of these enrichment cultures can O-demethylate MTBE leading to accumulation of tert-butyl alcohol (TBA).25,36 No subsequent degradation of TBA was observed in these cultures. Methane accumulation in the headspace indicated that methanogens may utilize the degradation products of MTBE.26 However, bromoethanesulfonic acid, a specific inhibitor of methanogenesis, did not inhibit MTBE degradation, indicating that the initial steps in biodegradation are mediated by bacteria.37 Analyses of anaerobic MTBE-degrading enrichment cultures and environmental samples, including SIP approaches, have revealed complex bacterial communities.38−41 However, identification of microorganisms involved in MTBE degradation were not always conclusive and hence, knowledge of the responsible microorganisms is still limited. In this study, we analyzed the bacterial community structure of a methanogenic enrichment culture from the AK in New Jersey and applied DNA-SIP to identify the bacteria active in assimilation of carbon from anaerobic biodegradation of MTBE. The SIP data indicated that phylotypes belonging to the Ruminococcaceae in the Firmicutes were active in the methanogenic cultures. A bacterium most closely related to Saccharofermentans acetigenes was shown to incorporate 13C-MTBE into its DNA, suggesting it is the principal microorganism active in O-demethylation of MTBE.

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MATERIALS AND METHODS Experimental Setup and Analytical Methods. Anaerobic enrichment cultures used in this study were originally established with 10% sediment (v/v) from the Arthur Kill Inlet (AK) between New Jersey and the New York Harbor.36 The enrichment cultures have been maintained for over a decade using strict anaerobic technique under methanogenic conditions at 28 °C.25,37 The cultures were repeatedly transferred at 3−6-month intervals and usually at a 1:5 to 1:10 dilution, into fresh medium and enriched with MTBE (Aldrich, Milwaukee, WI) as the sole carbon source.25,36,37 MTBE concentration was regularly monitored using gas chromatography with flame ionization detection.36 MTBE SIP Cultures. Six replicate cultures of the AK enrichment were set up for SIP experiments. These cultures were transferred to 60 mL glass serum vials, capped with Teflon-coated stoppers and aluminum crimp seals and incubated at 28 °C. Three experimental cultures were spiked with 100 μM MTBE with a 13C labeled O-methyl carbon (Campro Scientific GmbH, Berlin, Germany at >99% atom 13C purity; kindly provided by H. Richnow, UFZ, Leipzig). Triplicate control cultures were spiked with 100 μM 12CMTBE. Samples were taken for analysis at three time points after approximately 0, 20, and 50% MTBE depletion. Acetate and Cell Biomass SIP Cultures. On the basis of previous findings,37,38 we hypothesized that acetogenic bacteria are able to cleave the ether bond of MTBE and convert the 1456

DOI: 10.1021/acs.est.5b04731 Environ. Sci. Technol. 2016, 50, 1455−1464

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Inc., CA) in the dark for 5 min. Before mounting onto a slide with a drop of SlowFade Gold antifade reagent (Life Technologies, CA), the stained cells were filtered onto a 0.2 μm black polycarbonate filter paper and flushed with PBS to wash the sample. The cells were counted manually through epifluorescence images (ImageXpress, AXON, USA) at an excitation wavelength of 358 nm and an emission wavelength of 461 nm. Accession Numbers. The partial 16S rRNA gene sequences were deposited in Genbank under accession numbers KT279744-KT279757. Pyrosequencing data are available at BioProject ID PRJNA290405.

Approximately 15−20 ng of the PCR product was digested with restriction enzyme MnlI (New England BioLabs, MA) at 37 °C for 6 h. Digested samples were precipitated with sodium acetate and ethanol, resuspended in formamide with a ROX standard and denatured at 94 °C.42 Samples were analyzed on an ABI 310 Genetic analyzer with peak detection/quantification employing Genescan software (Applied Biosystems Instruments, Foster City, CA), which generated a fingerprint of each community by terminal restriction fragment length polymorphism (T-RFLP). Cloning, Sequencing, and Phylogenic Analysis. One 12 C MTBE-amended replicate microcosm after 50% degradation was selected for clone analysis. The 16S rRNA genes from this culture were amplified by PCR with 27F and 1100R primers and a clone library was generated using the Promega pGEM-T Vector System I cloning kit (Promega, Madison, WI). To identify the bacteria associated with corresponding terminal restriction fragments (T-RFs) observed in the enrichment cultures, each individual clone was screened by T-RFLP to match with predominant T-RFs in the community profile. Plasmid DNA of selected clones was extracted using Zyppy Plasmid Miniprep Kits (Zymo, Irvine, CA) and sequenced by Genewiz Inc. (South Plainfield, NJ). The T-RFs that composed 5% or more relative abundance in the communities were recovered from the clone library allowing us to identify the predominant bacterial phylotypes of the culture. Pyrosequencing. Triplicate MTBE-utilizing enrichment cultures were selected for pyrosequencing and community analysis. When 50% of amended 12C MTBE had been depleted, DNA was extracted from the triplicate cultures. 16S rRNA genes of the bacterial community were amplified with 27F and 519R primers, then sequenced and analyzed by Molecular Research LP (Shallowater, TX) using a Roche 454 Genome Sequencer following the manufacturer’s guidelines. The sequence data was processed using a proprietary analysis pipeline (MR DNA, Molecular Research LP). Briefly, the barcodes, primers, short sequences (6 bp) were removed to ensure the quality of reads. The remaining sequences were denoised, and then chimeras and singleton sequences were removed. Operational taxonomic units (OTUs) were clustered using a sequence similarity threshold of 97% and taxonomically classified by BLASTN against a curated GreenGenes database as described previously.47,48 A total of 12040, 7230, and 13371 quality-filtered reads, for each biological replicate, were grouped into 56 operational taxonomic units (OTUs) with 81.3% to 100% similarity, enabling identification of OTUs to the family level. Data Analysis. The 16S rRNA gene sequences were compared with the basic local alignment search tool (BLAST) and Ribosome Database Project (RDP) Seqmatch tool for taxonomic identification. Phylogenetic trees were constructed using the programs Geneious 8.1 (http://www.geneious. com)49 and MEGA 6.0.50 Clone sequences and selected sequences of the most closely related cultivated species determined from BLAST searches were used to construct a maximum likelihood tree.51 DAPI Staining. One Arthur Kill methanogenic 12C MTBE amended culture was selected for cell counting. One ml of cell culture was sonicated for 30 s and centrifuged at 2000g for 3 min to remove FeS precipitates in the medium. The cells were then stained with the dye 4′,6-diamidino-2-phenylindole dihydrochloride (DAPI, 12.5 μg mL−1, Vector Laboratories

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RESULTS Bacterial Community Analysis by DNA-SIP. The AK methanogenic MTBE-degrading culture, after 15 years of enrichment with MTBE as the sole carbon source, contained a simplified bacterial community consisting of approximately ten dominant phylotypes and multiple less abundant phylotypes (terminal restriction fragments; T-RFs) as represented in 16S rRNA gene terminal restriction fragment length polymorphism (T-RFLP) profiles (Supporting Information, Supplemental Figure 1). Over the course of the multiyear enrichment and transfer of the cultures there was some variation observed in peak intensity as well as the appearance and loss of some T-RFs (data not shown). DNA-SIP was applied to determine the initial bacterial phylotypes that obtained their cellular carbon from MTBE. Six replicate enrichment cultures were transferred 1:2.5 into fresh medium to yield a total dilution of 10−11 from the original Arthur Kill sediment. Triplicate cultures were amended with either 100 μM 13C-labeled MTBE or with 12C MTBE as the control. MTBE depletion occurred within 21 days in both the 13 C- and 12C-MTBE amended cultures with similar degradation rates (Figure 1). The concentration of TBA, the daughter product of MTBE degradation, increased corresponding with the decrease in MTBE concentration. No further degradation of TBA was observed in any of the enrichment cultures.

Figure 1. Anaerobic biodegradation of 12C-MTBE and 13C-MTBE to 12 C-TBA and 13C-TBA, respectively, by methanogenic Arthur Kill enrichment cultures. Error bars represent the mean and standard deviation of triplicate cultures. Arrows indicate sampling points for community analyses. 1457


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Figure 2. Control and experimental 16S rRNA gene stable isotope probing T-RFLP profiles from (blue) light 12C-DNA and (red) heavy 13C-DNA bands of MTBE-degrading methanogenic cultures. Replicate 1: (a) Time zero resident community; (b) 12C-MTBE-amended control after 20% degradation; and (c) 13C-MTBE-experimental SIP profiles after 20% degradation. Replicate 2: (d) Time zero resident community; (e) 12C-MTBEamended control after 20% degradation; and (f) 13C-MTBE-experimental SIP profiles after 20% degradation.

Figure 3. (a) Relative abundance of bacterial OTUs (%) identified in the AK culture clone libraries as determined by 16S rRNA gene T-RFLP. (b) Relative abundance of bacterial families identified as determined by 16S rRNA gene 454-pyrosequencing and clustered to the family level. Color coding is phylogenetically synchronized between T-RFLP and pyrosequencing data.

For each replicate culture, the SIP-T-RFLP fingerprints at different time points (Figure 2) illustrate the predominant OTUs of the active bacterial community (13C-DNA heavy band) compared to the resident community (12C-DNA light band). In time zero profiles, the predominant T-RFs were 95, 121, 167, 210, and 236 bp, accounting for 58 ± 4% of the total

T-RFLP peak area. We obtained sufficient DNA from two out of three 13C-MTBE treatment samples at time zero and 20% degradation to proceed to downstream analyses, but not from the samples from the 50% degradation time point. At 20% MTBE degradation, the predominant T-RFs of the resident communities were nearly identical between the 12C1458


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Figure 4. Maximum likelihood phylogenetic relationships of organisms detected in Arthur Kill MTBE-degrading methanogenic culture based on 16S rRNA gene sequences. We used 1073 bp unambiguously aligned nucleotide positions for analysis. Numbers at nodes indicate bootstrap values from 100 replications. Accession numbers for reference sequences are indicated in parentheses.

MTBE-control and 13C-MTBE-experimental treatments (Figure 2b,c; blue profile). Approximately 16−20 peaks were detected in the resident (12C-DNA light band) community profiles, dominated by T-RFs of 95, 121, 210, and 282 bp. The relative peak area of the 95, 121, and 210 bp T-RFs increased from 29 ± 6 to 45 ± 5% as MTBE was degraded (Figure 3a). No T-RFs were detected in the 13C-DNA heavy bands of 12CMTBE-fed controls (Figure 2b,e; red profile). In contrast, the 13 C-MTBE-fed experimental profiles (13C-DNA heavy band) contained a predominant 207 bp T-RF, suggesting that this OTU could initially utilize the methyl group of MTBE and assimilate this carbon into its biomass (Figure 2c,f; red profile). Identification of Active MTBE-Degrading Bacteria. To identify the active bacteria in MTBE-utilizing culture and assign phylogenetic affiliation to the various T-RFLP peaks, 16S rRNA gene clone libraries were constructed and screened to match specific T-RF peaks. Importantly, this approach yielded ∼1000 bp partial 16S rRNA gene sequences allowing for more detailed phylogenetic resolution of the community members. The clone

representing the 207 bp peak, the predominant 13C MTBE utilizing phylotype, clustered in the Firmicutes with Saccharofermentans acetigenes as its closest cultivated relative (94% 16S rRNA gene similarity over 1090 bp) and the closest match was to an uncultured clone from a PCB degrading community52 (99% similarity over 1069 bp). The other clones from the MTBE-utilizing culture spread across diverse phyla, including Synergistetes, Actinobacteria, Proteobacteria, and Firmicutes. The 93 bp T-RF and the other T-RFs with relatively low abundance were not recovered from the clone library. A maximum likelihood phylogenic tree of the clones and their closest cultured bacterial species was reconstructed (Figure 4). Many of the MTBE-culture clones were closely related to uncultured environmental sequences, but only distantly related to cultivated isolates. The pyrosequencing results agreed with the T-RFLP/clone library data. The dominant families were Desulfobulbaceae, Synergistaceae, Acidimicrobiales and Peptococcaceae which corresponded to the 210, 95, 121, and 222 bp T-RFs, respectively, 1459


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Figure 5. Control and experimental16S rRNA gene stable isotope probing T-RFLP profiles from (blue) light 12C-DNA and (red) heavy 13C-DNA bands of acetate amended cultures, demonstrating the resident and the active communities. (a) Four days incubation of 12C-acetate-amended control; (b) 4 days incubation of 13C-acetate experimental; (c) 1 month incubation of 12C-acetate-amended control; and (d) 1 month incubation of 13 C-acetate experimental. Specific TRFs of interest are indicated in the SIP profiles.

accounting for 60 ± 20% of the total reads (Supporting Information, Supplemental Table 1 and Supplemental Figure 2). This observation was in general agreement with the TRFLP/clone library results (50 ± 11% of total T-RF peak area). The relative abundance of the putative MTBE degrader, Ruminococcaceae, represented 3 ± 2% of the bacterial community in the pyrosequencing analysis which was similar to the T-RFLP peak area of 4 ± 2% (Figure 3b). Utilization of Acetate and Biomass. One byproduct of MTBE degradation is assumed to be acetate, which might be an important carbon source for the remainder of the enriched community. Figure 5 shows the results of a 13C-acetate SIP experiment from the AK enrichment culture. A 282 bp T-RF was the dominant 13C-labeled peak (Figure 5d; red profile). The closest isolated species corresponding to the 282 bp T-RF clone was Syntrophobacter f umaroxidans belonging to the Deltaproteobacteria.53,54 Interestingly, the closest match for the 282 bp T-RF was an uncultured clone from the same PCB degrading community as for the 207 bp T-RF (99% similarity over 1111 bp).52 The remaining T-RFs which persisted in the community profiles (i.e., 95, 210, 238, 273, 298, and 331 bp TRF peaks) are assumed to incorporate cellular biomass or other metabolites generated by T-RFs 207 and 282 rather than carbon directly from MTBE or acetate because these peaks were not immediately labeled with 13C-MTBE or 13C-acetate in our SIP experiments (Figure 5). To determine a possible role for these other persisting members of the MTBE-fed AK community, we examined the utilization of biomass using a DNA-SIP approach. The AK enrichment was spiked with 13C-labeled biomass (BioExpress Media). Several T-RF peaks (95, 210, 238, 273, 298, and 331 bp) appeared in the 13C-biomass amended experimental profile (Supporting Information, Supplemental Figure 3), indicating that the species representing these T-RFs likely utilized complex biomolecules as a primary carbon source, but not MTBE. Hence, although only the Ruminococcaceae phylotype

directly incorporated carbon from MTBE, the other members of the AK enrichment community were likely maintained as scavengers, cross feeding on the biomass and metabolites of the MTBE-utilizing acetogens and methanogens.

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DISCUSSION Anaerobic biodegradation of MTBE has been shown for a number of microcosms and field studies under different redox conditions.25,27−29,36 Our previous characterization of anaerobic MTBE-utilizing enrichment cultures demonstrated that Odemethylation of MTBE was likely mediated by acetogenic bacteria, with member of the Deltaproteobacteria, Chlorof lexi, and Firmicutes as the three dominant phylotypes in the community.37 Here, we provide evidence that the bacterium mediating the initial degradation of MTBE in a methanogenic enrichment culture is a member of the Ruminococcaceae. The 207 bp R-TF clone was most closely related to Saccharofermentans acetigenes, a newly discovered species within the Ruminococcaceae family, of the order Clostridiales in the Firmicutes, albeit at low (94%) 16S rRNA gene sequence similarity. S. acetigenes is an obligate anaerobic bacterium isolated from brewery wastewater treatment sludge.55 Only a few previous studies have applied the SIP technique in an attempt to identify microorganisms involved in MTBE degradation under anaerobic conditions, but the results were not always conclusive with multiple OTUs incorporating the 13 C-label.39−41 Of particular interest is that bacteria affiliated with Ruminococcaceae (JQ428827) and Sphingopyxis from were recognized as possible MTBE degraders in methanogenic MTBE-degrading microcosms established from activated sludge.40 Interestingly, the presence of a Ruminococcaceae (KC758914) had also been identified in another SIP study of MTBE contaminated groundwater in California.39 Our prior research indicated that O-demethylation is the initial step in cleavage of the ether bond of MTBE in the AK cultures.37 Furthermore, earlier characterization of the anaerobic MTBE1460


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of a portion of the acetogenic Ruminococcaceae population and detection by SIP. The SIP analysis of the acetate utilizing bacterial community suggested that acetate was not substantially converted to bacterial biomass and incorporated to DNA. Instead, we postulate that O-demethylating acetogens would feed acetoclastic methanogens, and then various fermenting organisms could slowly utilize the carbon from decaying biomass, originally derived from MTBE. Interestingly, several of the identified T-RFs in the MTBE-degrading community were closely related to sulfate-reducing bacteria. Their enrichment may be attributed to the fermentation of biomass or the utilization of trace amounts of sulfate in the media. After acetate amendment the relative abundance of the 131 bp T-RF in the MTBE-utilizing cultures substantially increased (Figure 5), suggesting the utilization of acetate. The closest species match for the 131 bp T-RF is Paenirhodobacter enshiensis (97% similarity over 1053 bp), a facultative anaerobic Alphaproteobacteria that grows chemoheterotrophically and can use acetate as a sole carbon source.68 The 167 bp T-RF clustered with Dechlorosoma suillum, which can also oxidize acetate.69 The 95, 121, and 210 bp T-RFs are distantly related to Thermanaerovibrio velox (89% similarity over 1082 bp), Gaiella occulta (86% similarity over 1100 bp) and Desulfobulbus propionicus (91% similarity over 1124 bp), respectively, and may consume secondary products of acetate and biomass (Supporting Information, Supplemental Figure 2). T. velox in the Synergistetes phylum is able to ferment a variety of organic substrates, but does not utilize acetate.70 Members of the Actinobacteria (95 bp T-RF) have also been detected in other anaerobic MTBE degrading cultures.40,71,72 The 210 bp T-RF clustered within D. propionicus, which has a broad versatility in using various fermentation.73 The 95 bp and 210 bp T-RFs were abundant in the 13C-Bioexpress amended community, indicating that they may utilize decayed biomass more efficiently than the species represented by either the 131 bp or 167 bp T-RFs (Supporting Information, Supplemental Figure 2). The 282 bp T-RF, most closely related to Syntrophobacter f umaroxidans (92% similarity over 1127 bp), was enriched in 13C-DNA heavy bands of the acetate-SIP cultures. Syntrophobacter species are often associated with degradation of propionate to acetate in methanogenic ecosystems.53,54 We conclude that all these phylotypes play secondary roles in the MTBE-degrading culture through the utilization of either acetate and biomass. In summary, acetogenic bacteria in the Ruminococcaceae are hypothesized to initially O-demethylate MTBE via the reductive CoA pathway due to their immediate incorporation of 13C from MTBE in our SIP experiments. Their activity is possibly facilitated by bacteria and methanogens that consume acetate. The other microorganisms in the methanogenic AK enrichment culture community likely metabolize the decayed biomass of acetogens and methanogens by fermentation. Our findings support the notion that Ruminococcaceae species are also the key bacteria in MTBE degradation that have been identified in the other SIP studies. A characterization of the metabolic pathways and identification of the microorganisms responsible for the activity will help us better understand anaerobic MTBE degradation processes in the field and determine biomarkers for monitoring natural attenuation.

degrading enrichment showed that propyl iodide, which binds corrinoids in a light-reversible manner, caused light-reversible inhibition of MTBE-degradation, suggesting that the MTBE degradation process was corrinoid-dependent.37 In addition, rifampicin, a bacterial protein synthesis inhibitor, prevented anaerobic MTBE degradation,37 while bromoethanesulfonate, an inhibitor of methanogenesis, did not block anaerobic MTBE utilization.25 Overall carbon flow in during MTBE degradation was eventually coupled to methanogenesis.25 Here, with the identification and phylogenic analysis of Ruminococcaceae as the putative MTBE-degrading bacteria in this study and other SIP experiments,39,40 it is reasonable to propose that acetogens belonging to Ruminococcaceae have the capability for Odemethylation and assimilation of the methyl carbon of MTBE. Many acetogenic bacteria are capable of utilizing methyl groups to synthesize acetate with formation of ATP via the reductive acetyl-coenzymeA pathway.56 In a typical reductive acetyl-CoA pathway, both carbon monoxide and methyl groups are reduced from carbon dioxide prior to generating acetate.57 However, the methyl group can also be O-demethylated from a range of natural methyl donors, such as ethers and aromatic compounds.58−61 The methyl group of MTBE is likely transferred to tetrahydrofuran (THF), a methyl-accepting corrinoid protein, and then to coenzyme A by a methyltransferase system, eventually yielding acetate.56,57,62 We previously demonstrated that anaerobic MTBE biodegradation by the AK cultures is a corrinoid-dependent process.37 Furthermore, methanogens in the community25 might assist anaerobic MTBE degradation via the reversed Wood-Ljungdahl pathway as a result of oxidation of acetate to methane and releasing two electrons to drive the reductive processes of the acetyl-CoA pathway.62 Because our focus in this study was to specifically identify the bacteria mediating the initial attack, we did not attempt to identify any methanogens that would utilize the MTBE degradation products. It should be noted that the theoretical energy available for growth from acetogenesis may not directly translate to cell yield.63 When CO is utilized as the substrate, the acetate-tobiomass ratio of Clostridium thermoaceticum is 60,64,65 indicating that carbon assimilated into biomass of acetogenic bacteria, and hence incorporation of labeled carbon into DNA in a SIP experiment, can be limited. The estimated initial cell concentration of the AK methanogenic culture used for the SIP experiments in this study was 4 × 105 cells ml−1 from DAPI staining (data not shown) and the relative abundance of the 207 bp T-RF was 2−5% of the bacterial community (Figures 2 and 3). From the rate of MTBE loss (Figure 1) and bacterial community abundance, the estimated MTBE degradation rate was 2−5 × 10−5 μg MTBE Ruminococcaceae cell−1 day−1. This estimation was based on the assumptions that the total cell density remained essentially constant over the incubation period, and MTBE was only utilized by Ruminococcaceae. From the cell density of the enrichment (4 × 105 cells ml−1), the relative abundance of Ruminococcaceae (5%), and assuming the archaeal population was below 10% of the total cell numbers, we can estimate that there are ∼2 × 104 ml−1 Ruminococcaceae cells within the enrichment. Given an estimated dry weight of 3 × 10−13 g for a bacterial cell and 50% C content,66,67 this would indicate a cellular dry weight of ∼1× 10−7 g mL−1, corresponding to ∼6× 10−8 g mL−1 C. On the basis of this estimation, the methyl carbon from 20 μmol l−1 of 13C-MTBE (that had been utilized at the SIP sampling time point) provided sufficient carbon (24× 10−8 g mL−1 C) for the growth 1461


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(12) Ayotte, J. D.; Argue, D. M.; McGarry, F. J. Methyl tert-butyl ether occurrence and related factors in public and private wells in southeast New Hampshire. Environ. Sci. Technol. 2005, 39 (1), 9−16. (13) van Wezel, A.; Puijker, L.; Vink, C.; Versteegh, A.; de Voogt, P. Odour and flavour thresholds of gasoline additives (MTBE, ETBE and TAME) and their occurrence in Dutch drinking water collection areas. Chemosphere 2009, 76, 672−676. (14) Juhler, R. K.; Felding, G. Monitoring methyl tertiary butyl ether (MTBE) and other organic micropollutants in groundwater: results from the Danish National Monitoring Program. Water, Air, Soil Pollut. 2003, 149 (1−4), 145−161. (15) Squillace, P. J.; Moran, M. J.; Price, C. V. VOCs in shallow groundwater in new residential/commercial areas of the United States. Environ. Sci. Technol. 2004, 38 (20), 5327−5338. (16) McMahon, P.; Chapelle, F. Redox processes and water quality of selected principal aquifer systems. Groundwater 2008, 46 (2), 259− 271. (17) Häggblom, M. M.; Youngster, L. K.; Somsamak, P.; Richnow, H. H. Anaerobic biodegradation of methyl tert-butyl ether (MTBE) and related fuel oxygenates. Adv. Appl. Microbiol. 2007, 62 (62), 1−20. (18) Keller, A. A.; Sandall, O. C.; Rinker, R. G.; Mitani, M. M.; Bierwagen, B.; Snodgrass, M. J. Cost and performance evaluation of treatment technologies for MTBE-contaminated water. UC TSR&TP Report to the Governor of California, 1998. (19) Van Deuren, J.; Lloyd, T.; Chhetry, S.; Liou, R.; Peck, J. 3.9 In situ biological treatment for ground water, surface water, and leachate, remediation technologies screening matrix and reference guide, Version 4th ed.. Federal Remediation Technologies Roundtable, PII-0111641. (20) USEPA. Monitored Natural Attenuation of MTBE as a Risk Management Option at Leaking Underground Storage Tank Sites. Office of Research and Development National Risk Management Research Laboratory, EPA600/R-04/179. (21) Sarkar, D.; Ferguson, M.; Datta, R.; Birnbaum, S. Bioremediation of petroleum hydrocarbons in contaminated soils: comparison of biosolids addition, carbon supplementation, and monitored natural attenuation. Environ. Pollut. 2005, 136 (1), 187− 195. (22) Suflita, J. M.; Mormile, M. R. Anaerobic biodegradation of known and potential gasoline oxygenates in the terrestrial subsurface. Environ. Sci. Technol. 1993, 27 (5), 976−978. (23) Mormile, M. R.; Liu, S.; Suflita, J. M. Anaerobic biodegradation of gasoline oxygenates: extrapolation of information to multiple sites and redox conditions. Environ. Sci. Technol. 1994, 28 (9), 1727−1732. (24) Bradley, P. M.; Chapelle, F. H.; Landmeyer, J. E. Effect of redox conditions on MTBE biodegradation in surface water sediments. Environ. Sci. Technol. 2001, 35 (23), 4643−4647. (25) Somsamak, P.; Richnow, H. H.; Häggblom, M. M. Carbon isotopic fractionation during anaerobic biotransformation of methyl tert-butyl ether and tert-amyl methyl ether. Environ. Sci. Technol. 2005, 39 (1), 103−109. (26) Somsamak, P.; Richnow, H. H.; Häggblom, M. M. Carbon isotope fractionation during anaerobic degradation of methyl tert-butyl ether under sulfate-reducing and methanogenic conditions. Appl. Environ. Microbiol. 2006, 72 (2), 1157−1163. (27) Finneran, K. T.; Lovley, D. R. Anaerobic degradation of methyl tert-butyl ether (MTBE) and tert-butyl alcohol (TBA). Environ. Sci. Technol. 2001, 35 (9), 1785−1790. (28) Wilson, J. T.; Adair, C.; Kaiser, P. M.; Kolhatkar, R. Anaerobic biodegradation of MTBE at a gasoline spill site. Groundwater Monit. Rem. 2005, 25 (3), 103−115. (29) Pruden, A.; Sedran, M. A.; Suidan, M. T.; Venosa, A. D. Anaerobic biodegradation of methyl tert-butyl ether under ironreducing conditions in batch and continuous-flow cultures. Water Environ. Res. 2005, 77, 297−303. (30) Fischer, A.; Oehm, C.; Selle, M.; Werner, P. Biotic and abiotic transformations of methyl tertiary butyl ether (MTBE). Environ. Sci. Pollut. Res. 2005, 12 (6), 381−386. (31) Wellington, E. M.; Berry, A.; Krsek, M. Resolving functional diversity in relation to microbial community structure in soil:

ASSOCIATED CONTENT

S Supporting Information *

The Supporting Information is available free of charge on the ACS Publications website at DOI: 10.1021/acs.est.5b04731. Three additional figures and one table as noted in text. (PDF)

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AUTHOR INFORMATION

Corresponding Author

*E-mail: [email protected]. Tel.: (+1) 848 932 5646. Fax: (+1) 848 932 8965. Notes

The authors declare no competing financial interest.

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ACKNOWLEDGMENTS We thank Dr. H−H Richnow, Helmholtz Centre for Environmental Research - UFZ Leipzig for the gift of 13C labeled MTBE. This work was supported in part by the National Science Foundation (CEBT 1335824), the New Jersey Water Resources Research Institute, and the New Jersey Agricultural Experiment Station.

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REFERENCES

(1) Deeb, R. A.; Scow, K. M.; Alvarez-Cohen, L. Aerobic MTBE biodegradation: an examination of past studies, current challenges and future research directions. Biodegradation 2000, 11 (2−3), 171−185. (2) Franklin, P. M.; Koshland, C. P.; Lucas, D.; Sawyer, R. F. Clearing the air: Using scientific information to regulate reformulated fuels. Environ. Sci. Technol. 2000, 34 (18), 3857−3863. (3) Squillace, P. J.; Moran, M. J. Factors associated with sources, transport, and fate of volatile organic compounds and their mixtures in aquifers of the United States. Environ. Sci. Technol. 2007, 41 (7), 2123−2130. (4) AWWA. Two updated MTBE contamination cost analyses pin MTBE clean up costs between $25−$85 billion. American Water Works Association: Denver, CO, 2005. (5) Carter, J. M.; Grady, S. J.; Delzer, G. C.; Koch, B.; Zogorski, J. S. Occurrence of MTBE and other gasoline oxygenates in CWS source waters. J. Am. Water. Resour. As. 2006, 98 (4), 91−104. (6) Sweet, F.; Kauffman, M.; Pellerin, T.; Espy, D.; Mills, M. An estimate of the national cost for remediation of MTBE releases from existing leaking underground storage tank sites. Proceedings of the Annual International Conference on Soils, Sediments, Water and Energy 2010, 11 (1), 12. (7) Zogorski, J. S.; Carter, J. M.; Ivahnenko, T.; Lapham, W. W.; Moran, M. J.; Rowe, B. L.; Squillace, P. J.; Toccalino, P. L. Volatile organic compounds in the nation’s ground water and drinking-water supply wells. U.S. Geological Survey Circular 2006, 1292, 101. (8) Shamsipur, M.; Miran Beigi, A.; Teymouri, M.; Poursaberi, T.; Mostafavi, S. M.; Soleimani, P.; Chitsazian, F.; Tash, S. Biotransformation of methyl tert-butyl ether by human cytochrome P450 2A6. Biodegradation 2012, 23 (2), 311−318. (9) Bermudez, E.; Willson, G.; Parkinson, H.; Dodd, D. Toxicity of methyl tertiary-butyl ether (MTBE) following exposure of Wistar Rats for 13 weeks or one year via drinking water. J. Appl. Toxicol. 2012, 32 (9), 687−706. (10) Stocking, A. J.; Deeb, R. A.; Flores, A. E.; Stringfellow, W.; Talley, J.; Brownell, R.; Kavanaugh, M. C. Bioremediation of MTBE: a review from a practical perspective. Biodegradation 2000, 11 (2−3), 187−201. (11) Deeb, R. A.; Chu, K.; Shih, T.; Linder, S.; Suffet, I.; Kavanaugh, M. C.; Alvarez-Cohen, L. MTBE and other oxygenates: environmental sources, analysis, occurrence, and treatment. Environ. Eng. Sci. 2003, 20 (5), 433−447. 1462


Article

Environmental Science & Technology exploiting genomics and stable isotope probing. Curr. Opin. Microbiol. 2003, 6 (3), 295−301. (32) Dumont, M. G.; Murrell, J. C. Stable isotope probinglinking microbial identity to function. Nat. Rev. Microbiol. 2005, 3 (6), 499− 504. (33) Madsen, E. L. The use of stable isotope probing techniques in bioreactor and field studies on bioremediation. Curr. Opin. Biotechnol. 2006, 17 (1), 92−97. (34) Kerkhof, L. J. and Häggblom, M. M. Detecting active bacteria involved in biodegradation in the environment, In Microbial Biodegradation: Genomics and Molecular Biology, Díaz, E., Ed.; Caister Academic Press: Norfolk, England, 2008; pp 55−70. (35) Radajewski, S.; Ineson, P.; Parekh, N. R.; Murrell, J. C. Stableisotope probing as a tool in microbial ecology. Nature 2000, 403 (6770), 646−649. (36) Somsamak, P.; Cowan, R. M.; Häggblom, M. M. Anaerobic biotransformation of fuel oxygenates under sulfate-reducing conditions. FEMS Microbiol. Ecol. 2001, 37 (3), 259−264. (37) Youngster, L.; Somsamak, P.; Häggblom, M. Effects of cosubstrates and inhibitors on the anaerobic O-demethylation of methyl tert-butyl ether (MTBE). Appl. Microbiol. Biotechnol. 2008, 80 (6), 1113−1120. (38) Youngster, L. K. G.; Kerkhof, L. J.; Häggblom, M. M. Community characterization of anaerobic methyl tert-butyl ether (MTBE)-degrading enrichment cultures. FEMS Microbiol. Ecol. 2010, 72 (2), 279−288. (39) Key, K. C.; Sublette, K. L.; Duncan, K.; Mackay, D. M.; Scow, K. M.; Ogles, D. Using DNA-stable isotope probing to identify MTBEand TBA-degrading microorganisms in contaminated groundwater. Ground Water Monit. & Remed. 2013, 33 (4), 57−68. (40) Sun, W.; Sun, X.; Cupples, A. M. Anaerobic methyl tert-butyl ether-degrading microorganisms identified in wastewater treatment plant samples by stable isotope probing. Appl. Environ. Microbiol. 2012, 78 (8), 2973−2980. (41) Busch-Harris, J.; Sublette, K.; Roberts, K. P.; Landrum, C.; Peacock, A. D.; Davis, G.; Ogles, D.; Holmes, W. E.; Harris, D.; Ota, C.; Yang, X.; Kolhatkar, A. Bio-traps coupled with molecular biological methods and stable isotope probing demonstrate the in situ biodegradation potential of MTBE and TBA in gasoline-contaminated aquifers. Groundwater Monit. Rem. 2008, 28 (4), 47−62. (42) Gallagher, E.; McGuinness, L.; Phelps, C.; Young, L. Y.; Kerkhof, L. J. 13C-carrier DNA shortens the incubation time needed to detect benzoate-utilizing denitrifying bacteria by stable-isotope probing. Appl. Environ. Microbiol. 2005, 71 (9), 5192−5196. (43) Kerkhof, L.; Ward, B. B. Comparison of nucleic acid hybridization and fluorometry for measurement of the relationship between RNA/DNA ratio and growth rate in a marine bacterium. Appl. Environ. Microbiol. 1993, 59 (5), 1303−1309. (44) Sakano, Y.; Kerkhof, L. Assessment of changes in microbial community structure during operation of an ammonia biofilter with molecular tools. Appl. Environ. Microbiol. 1998, 64 (12), 4877−4882. (45) Kerkhof, L. J.; Williams, K. H.; Long, P. E.; McGuinness, L. R. Phase preference by active, acetate-utilizing bacteria at the Rifle, CO integrated field research challenge site. Environ. Sci. Technol. 2011, 45 (4), 1250−1256. (46) Tuorto, S. J.; Darias, P.; McGuinness, L. R.; Panikov, N.; Zhang, T.; Häggblom, M. M.; Kerkhof, L. J. Bacterial genome replication at subzero temperatures in permafrost. ISME J. 2014, 8 (1), 139−149. (47) Dowd, S. E.; Sun, Y.; Secor, P. R.; Rhoads, D. D.; Wolcott, B. M.; James, G. A.; Wolcott, R. D. Survey of bacterial diversity in chronic wounds using pyrosequencing, DGGE, and full ribosome shotgun sequencing. BMC Microbiol. 2008, 8, 43. (48) Dowd, S.; Callaway, T.; Wolcott, R.; Sun, Y.; McKeehan, T.; Hagevoort, R.; Edrington, T. Evaluation of the bacterial diversity in the feces of cattle using 16S rDNA bacterial tag-encoded FLX amplicon pyrosequencing (bTEFAP). BMC Microbiol. 2008, 8, 125. (49) Kearse, M.; Moir, R.; Wilson, A.; Stones-Havas, S.; Cheung, M.; Sturrock, S.; Buxton, S.; Cooper, A.; Markowitz, S.; Duran, C.; Thierer, T.; Ashton, B.; Meintjes, P.; Drummond, A. Geneious Basic: an

integrated and extendable desktop software platform for the organization and analysis of sequence data. Bioinformatics 2012, 28 (12), 1647−1649. (50) Tamura, K.; Stecher, G.; Peterson, D.; Filipski, A.; Kumar, S. MEGA6: Molecular evolutionary genetics analysis version 6.0. Mol. Biol. Evol. 2013, 30 (12), 2725−2729. (51) Guindon, S.; Gascuel, O. A simple, fast, and accurate algorithm to estimate large phylogenies by maximum likelihood. Syst. Biol. 2003, 52 (5), 696−704. (52) Ho, C. H.; Liu, S. M. Impact of coplanar PCBs on microbial communities in anaerobic estuarine sediments. J. Environ. Sci. Health, Part B 2010, 45 (5), 437−448. (53) Harmsen, H. J.; Van Kuijk, B. L.; Plugge, C. M.; Akkermans, A. D.; De Vos, W. M.; Stams, A. J. Syntrophobacter f umaroxidans sp. nov., a syntrophic propionate-degrading sulfate-reducing bacterium. Int. J. Syst. Bacteriol. 1998, 48, 1383−1387. (54) Plugge, C. M.; Henstra, A. M.; Worm, P.; Swarts, D. C.; Paulitsch-Fuchs, A. H.; Scholten, J. C.; Lykidis, A.; Lapidus, A. L.; Goltsman, E.; Kim, E.; et al. Complete genome sequence of Syntrophobacter f umaroxidans strain (MPOBT). Stand. Genomic Sci. 2012, 7 (1), 91. (55) Chen, S.; Niu, L.; Zhang, Y. Saccharofermentans acetigenes gen. nov., sp. nov., an anaerobic bacterium isolated from sludge treating brewery wastewater. Int. J. Syst. Evol. Microbiol. 2010, 60 (12), 2735− 2738. (56) Fuchs, G. Alternative pathways of carbon dioxide fixation: insights into the early evolution of life? Annu. Rev. Microbiol. 2011, 65, 631−658. (57) Drake, H. L.; Küsel, K.; Matthies, C. Acetogenic Prokaryotes, In The Prokaryotes; Dworkin, M., et al. eds. Springer: New York, 2006; pp 354−420. (58) Emde, R.; Schink, B. Fermentation of triacetin and glycerol by Acetobacterium sp. No energy is conserved by acetate excretion. Arch. Microbiol. 1987, 149 (2), 142−148. (59) Wagener, S.; Schink, B. Fermentative degradation of nonionic surfactants and polyethylene glycol by enrichment cultures and by pure cultures of homoacetogenic and propionate-forming bacteria. Appl. Environ. Microbiol. 1988, 54 (2), 561−565. (60) Tanaka, K.; Pfennig, N. Fermentation of 2-methoxyethanol by Acetobacterium malicum sp. nov. and Pelobacter venetianus. Arch. Microbiol. 1988, 149 (3), 181−187. (61) Schramm, E.; Schink, B. Ether-cleaving enzyme and diol dehydratase involved in anaerobic polyethylene glycol degradation by a new Acetobacterium sp. Biodegradation 1991, 2 (2), 71−79. (62) Ragsdale, S. W.; Pierce, E. Acetogenesis and the Wood− Ljungdahl pathway of CO2 fixation. Biochim. Biophys. Acta, Proteins Proteomics 2008, 1784 (12), 1873−1898. (63) Drake, H. L. Acetogenesis, acetogenic bacteria, and the acetylCoA “Wood/Ljungdahl” pathway: past and current perspectives, In Acetogenesis; Drake, H. L., Ed.; Springer: US, 1994; pp 3−60. (64) Daniel, S. L.; Hsu, T.; Dean, S. I.; Drake, H. L. Characterization of the H2- and CO-dependent chemolithotrophic potentials of the acetogens Clostridium thermoaceticum and Acetogenium kivui. J. Bacteriol. 1990, 172 (8), 4464−4471. (65) Daniel, S. L.; Drake, H. L. Oxalate- and glyoxylate-dependent growth and acetogenesis by Clostridium thermoaceticum. Appl. Environ. Microbiol. 1993, 59 (9), 3062−3069. (66) Neidhardt, F. C. Chemical composition of Escherichia coli. In: Neidhardt, F. C., Ingraham, J. L., et al., editors. Escherichia coli and Salmonella: Cellular and Molecular Biology. ASM Press: Washington, D.C., 1987. (67) Taylor, A. R. Chemical analysis of the T2 bacteriophage and its host. Escherichia coli strain B. J. Biol. Chem. 1946, 165, 271−284. (68) Wang, D.; Liu, H.; Zheng, S.; Wang, G. Paenirhodobacter enshiensis gen. nov., sp. nov., a non-photosynthetic bacterium isolated from soil, and emended descriptions of the genera Rhodobacter and Haematobacter. Int. J. Syst. Evol. Microbiol. 2014, 64 (2), 551−558. (69) Achenbach, L. A.; Michaelidou, U.; Bruce, R. A.; Fryman, J.; Coates, J. D. Dechloromonas agitata gen. nov., sp. nov. and 1463


Article

Environmental Science & Technology Dechlorosoma suillum gen. nov., sp. nov., two novel environmentally dominant (per)chlorate-reducing bacteria and their phylogenetic position. Int. J. Syst. Evol. Microbiol. 2001, 51 (2), 527−533. (70) Zavarzina, D. G.; Zhilina, T. N.; Tourova, T. P.; Kuznetsov, B. B.; Kostrikina, N. A.; Bonch-Osmolovskaya, E. A. Thermanaerovibrio velox sp. nov., a new anaerobic, thermophilic, organotrophic bacterium that reduces elemental sulfur, and emended description of the genus Thermanaerovibrio. Int. J. Syst. Evol. Microbiol. 2000, 50 (3), 1287− 1295. (71) Wei, N.; Finneran, K. Microbial community analyses of three distinct, liquid cultures that degrade methyl tert-butyl ether using anaerobic metabolism. Biodegradation 2009, 20 (5), 695−707. (72) Raynal, M.; Crimi, B.; Pruden, A. Enrichment and characterization of MTBE-degrading cultures under iron and sulfate reducing conditions. Can. J. Civ. Eng. 2010, 37 (4), 522−534. (73) Pagani, I.; Lapidus, A.; Nolan, M.; Lucas, S.; Hammon, N.; Deshpande, S.; Cheng, J. F.; Chertkov, O.; Davenport, K.; Tapia, R.; Han, C.; Goodwin, L.; Pitluck, S.; Liolios, K.; Mavromatis, K.; Ivanova, N.; Mikhailova, N.; Pati, A.; Chen, A.; Palaniappan, K.; Land, M.; Hauser, L.; Chang, Y. J.; Jeffries, C. D.; Detter, J. C.; Brambilla, E.; Kannan, K. P.; Djao, O. D.; Rohde, M.; Pukall, R.; Spring, S.; Goker, M.; Sikorski, J.; Woyke, T.; Bristow, J.; Eisen, J. A.; Markowitz, V.; Hugenholtz, P.; Kyrpides, N. C.; Klenk, H. P. Complete genome sequence of Desulfobulbus propionicus type strain (1pr3). Stand. Genomic Sci. 2011, 4 (1), 100−110.

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