Cryogenic soil coring reveals coexistence of aerobic and anaerobic vinyl chloride degrading bacteria in a chlorinated ethene contaminated aquifer

Water Research 157 (2019) 281e291

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Cryogenic soil coring reveals coexistence of aerobic and anaerobic vinyl chloride degrading bacteria in a chlorinated ethene contaminated aquifer Patrick M. Richards a, 1, Yi Liang b, 1, Richard L. Johnson c, Timothy E. Mattes a, * a b c

Department of Civil and Environmental Engineering, 4105 Seamans Center, The University of Iowa, Iowa City, IA, 52242, USA State Key Laboratory of Organic Geochemistry, Guangzhou Institute of Geochemistry, Chinese Academy of Sciences, Guangzhou, 510640, China OHSU-PSU School of Public Health, Oregon Health & Science University, 3181 SW Sam Jackson Park Road, Portland, OR, 97239, USA

a r t i c l e i n f o

a b s t r a c t

Article history: Received 13 December 2018 Received in revised form 4 March 2019 Accepted 26 March 2019 Available online 28 March 2019

Vinyl chloride (VC) is a common groundwater contaminant and known human carcinogen. Three major bacterial guilds are known to participate in VC biodegradation: aerobic etheneotrophs and methanotrophs, and anaerobic organohalide-respiring VC-dechlorinators. We investigated the spatial relationships between functional genes representing these three groups of bacteria (as determined by qPCR) with chlorinated ethene concentrations in a surficial aquifer at a contaminated site. We used cryogenic soil coring to collect high-resolution aquifer sediment samples and to preserve sample geochemistry and nucleic acids under field conditions. All samples appeared to be anaerobic (i.e., contained little to no dissolved oxygen). VC biodegradation associated functional genes from etheneotrophs (etnC and/or etnE), methanotrophs (mmoX and/or pmoA), and anaerobic VC-dechlorinators (bvcA and/or vcrA) coexisted in 48% of the samples. Transcripts of etnC/etnE and bvcA/vcrA were quantified in contemporaneous groundwater samples, indicating co-located gene expression. Functional genes from etheneotrophs and anaerobic VC-dechlorinators were correlated to VC concentrations in the lower surficial aquifer (p < 0.05). Methanotroph functional genes were not correlated to VC concentrations. Cryogenic soil coring proved to be a powerful tool for capturing high-spatial resolution trends in geochemical and nucleic acid data in aquifer sediments. We conclude that both aerobic etheneotrophs and anaerobic VCdechlorinators may play a significant role in VC biodegradation in aquifers that have little dissolved oxygen. © 2019 Elsevier Ltd. All rights reserved.

Keywords: Vinyl chloride Etheneotrophs Methanotrophs Anaerobic VC-Dechlorinator Cryogenic coring Chlorinated solvents Aquifer sediments

1. Introduction The chlorinated solvents tetrachloroethene (PCE) and trichloroethene (TCE) have wide industrial application, primarily as dry cleaning fluids and degreasers (Watts, 1998). Poor handling practices historically have resulted in widespread groundwater contamination by these chemicals (Bradley, 2003; Moran et al., 2007). PCE and TCE are biodegradable by anaerobic bacteria through reductive dechlorination, which generates the toxic intermediates cis-dichloroethene (cis-DCE) and vinyl chloride (VC), a known human carcinogen (Clewell et al., 2001). Complete dechlorination of VC to environmentally-innocuous ethene is mediated by

* Corresponding author. E-mail address: [email protected] (T.E. Mattes). 1 These authors contributed equally to this work. https://doi.org/10.1016/j.watres.2019.03.059 0043-1354/© 2019 Elsevier Ltd. All rights reserved.

certain obligate organohalide-respiring bacteria (hereafter collectively referred to as anaerobic VC-dechlorinators). This includes Dehalococcoides mccartyii strains that express the reductive dehalogenase genes bvcA or vcrA (Krajmalnik-Brown et al., 2004; Müller et al., 2004) or the newly discovered Dehalogenimonas strain that expresses related dehalogenases (Yang et al., 2017). However, despite the presence of these bacteria at contaminated sites reductive dechlorination may frequently “stall”, resulting in the accumulation of cis-DCE and VC (Cox, 2012). Certain aerobic bacteria, including ethene-oxidizers (etheneotrophs) and methane-oxidizers (methanotrophs), are also known to effectively degrade VC into non-toxic end products (Coleman et al., 2002; Davis and Carpenter, 1990; Jin and Mattes, 2010; Verce et al., 2000). VC oxidation occurs by co-metabolism, in which methane or ethene are consumed as primary substrates, or through metabolic consumption by some etheneotrophs (VC-assimilators)(Fogel et al.,

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1986; Freedman and Herz, 1996; Jin and Mattes, 2008; Verce et al., 2001). Methanotrophs and etheneotrophs both express di-iron monooxygenases that can catalyze VC oxidation. During growth on ethene and VC, etheneotrophs express a soluble alkene monooxygenase (AkMO), transforming VC to the epoxide chlorooxirane, which is further metabolized to 2-chloro-2-hydroxyethyl-CoM by an epoxyalkane:coenzyme M transferase (EaCoMT)(Coleman and Spain, 2003). The genes etnC and etnE encode the a-subunit of the AkMO and the EaCoMT, respectively, and serve as specific biomarkers for etheneotrophs (Coleman and Spain, 2003; Liu et al., 2018). Methanotrophs express both soluble and particulate (i.e., membrane-bound) methane monooxygenases during growth on methane. The genes mmoX and pmoA encode the a-subunits of the soluble and particulate monooxygenases, respectively, and serve as specific biomarkers for methanotrophs (Paszczynski et al., 2011). Both metabolic and co-metabolic VC oxidation may be environmentally-relevant, since methane and ethene are often present in VC plumes (Findlay et al., 2016; Liang et al., 2017b). Aerobic VC biodegradation processes can be overlooked because the groundwater plumes where VC occurs are generally considered anaerobic. However, VC-oxidizing bacteria have been isolated from anaerobic groundwater (Fullerton et al., 2014), and laboratory studies have shown that pure cultures of certain etheneotrophs degrade VC at very low oxygen concentrations (Coleman et al., 2002). VC is more mobile than higher chlorinated ethenes and thus has greater potential to migrate away from anaerobic zones and into aerobic or micro-aerobic regions of a groundwater plume. Microcosm studies have demonstrated continued VC oxidation at oxygen concentrations below the detection limit of most field oxygen probes (0.2 mg/L) when a known low oxygen flux was provided (Gossett, 2010). It has been suggested that aerobic microorganisms are active in groundwater plumes that would be deemed anaerobic by most field measurements (Gossett, 2010). A recent study of functional gene expression by etheneotrophs, methanotrophs, and anaerobic VC-dechlorinators in 95 groundwater samples collected from monitoring wells across six chlorinated ethene contaminated sites found that etheneotrophs and methanotrophs were present in all of the samples (Liang et al., 2017b). In addition, mRNA transcripts of these functional genes were measured, which indicated that these organisms were active in the majority (66e86%) of the samples, even where the groundwater appeared anaerobic (Liang et al., 2017b). Surprisingly, these biomarkers often coincided with biomarkers for the anaerobic Dehalococcoides (Liang et al., 2017b), which are considered to be very sensitive to oxygen (Amos et al., 2008). Finding both aerobic and anaerobic bacteria in the same groundwater sample is unexpected, and leaves uncertainty about the source of these genes. Groundwater samples from typical monitoring wells are inherently composited from an undefined volume with potentially different biogeochemical conditions surrounding the well screen. Multilevel sampler data from a well-characterized site has shown that chlorinated ethene concentrations can vary by orders of magnitude within 1 m vertically (i.e., normal to groundwater flow), and that heterogeneity may contribute significantly to this variability near source areas (McMillan et al., 2018; Rivett et al., 2014). Comparison of multilevel sampler data with adjacent conventional monitoring wells shows that the monitoring well data is biased towards high permeability zones (McMillan et al., 2018). Aquifer core samples can provide discrete data from known locations within a plume, but can be difficult to retrieve under field conditions (Kiaalhosseini et al., 2016). In addition, biomarker and chemical data may be sensitive to environmental changes (i.e., changes in saturation or redox potential) that occur during drilling and sample retrieval (Kiaalhosseini et al., 2016). A newly developed cryogenic sampling technique freezes samples in situ, thus

preserving biomarkers and chemical analytes under field conditions (Kiaalhosseini et al., 2016), allowing for the analysis of biomarkers and chemical data in the subsurface at a high spatial resolution. The primary goal of this study is to use cryogenic coring to demonstrate the distribution of contaminants and VC biodegradation biomarker genes under very-high-resolution conditions where geochemical conditions are rigorously preserved due to in situ cryogenic freezing. Applying qPCR methods to these samples, we determined the distribution of functional genes from aerobic VC-oxidizers (i.e., methanotrophs and etheneotrophs) and anaerobic VC-dechlorinating bacteria within the aquifer at fine spatial resolution. We also measured geochemical parameters at this fine resolution to explore relationships between the abundance of VCdegrading bacterial functional genes and geochemical conditions in the samples. 2. Materials and methods 2.1. Site description MCRD Parris Island is a military training site located along the southern coast of South Carolina. At MCRD Parris Island there is a former dry cleaning facility, known as Site 45, experienced multiple PCE spills in the 1990s (Churchill, 2012; Vroblesky et al., 2009). Prior investigations at the site have delineated the contaminant plumes, and found evidence of significant PCE biodegradation including the generation of VC and other dechlorination intermediates (Churchill, 2012; Vroblesky et al., 2009). Several remediation pilot studies have been performed at the site, including one using emulsified zero-valent iron in the northern plume source area (Su et al., 2012). 2.2. Groundwater and cryo-core collection procedures Soil and groundwater samples were collected from the site in June 2016. Groundwater samples were collected using a previously described low-flow procedure (Liang et al., 2017b). Typical groundwater parameters were measured in the field during sample collection (Liang et al., 2017b). Groundwater samples were passed through Sterivex-GP filters (pore size 0.22 mM, Millipore, Germany) for duplicate DNA and RNA extractions (up to 1 L for DNA extraction, and up to 3 L for RNA extraction). Four soil cores were collected along a transect parallel to the groundwater flow; starting in the source area (Fig. 1). Soil cores were collected using a hollowstem auger modified for in situ cryogenic freezing with liquid nitrogen, as previously described (Kiaalhosseini et al., 2016). Frozen aquifer sediment cores were collected in sections of approximately 1.2 m, to a maximum depth of 5.5 m below ground surface (bgs). Each section was returned to the surface and subdivided. Top soil (0e0.6 m bgs) was discarded, and a miter saw was used to cut a 2.5 cm thick section every 15 cm in core depth, resulting in approximately 35 samples per core. Some of the sections had incomplete recovery because of blockages and soil loss from the bottom of the cores during retrieval; a total of 124 samples were collected. The samples were wrapped in aluminum foil, sealed in plastic bags, and shipped to the laboratory on dry ice. Samples were stored at
80  C until analyzed. 2.3. Cryo-core sample processing Frozen soil samples were processed based on previously described methods (Olson et al., 2017; Sale et al., 2015). Each sample was split into quarter sections using a hammer and masonry chisel and subjected to extraction procedures for chemicals

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Fig. 1. Aerial view of Site 45. The source area is outlined in red and the approximate extent of the northern plume (Vroblesky et al., 2009) is outlined in green. Monitoring wells nearby the four cryo-cores are indicated in blue. (For interpretation of the references to colour in this figure legend, the reader is referred to the Web version of this article.)

and nucleic acids described below. No appreciable thawing was noted during sample processing. All tools and surfaces were disinfected with 70% ethanol between samples. Samples were visually inspected during processing, and physical characteristics of the soil were noted. 2.4. Chemical extraction Aqueous extracts were prepared for geochemical analysis detailed below by filling pre-weighed 4 oz (140 ml) septa jars (Environmental Sampling Supply, Leandro, CA) with de-aired, deionized water. A frozen soil sample was added to each jar, allowing water to overflow, and capped with no headspace. The weight of the filled jars and the overflow were recorded and used to determine soil mass. The remaining water volume was also determined and used to calculate analyte concentrations in the soil. Methanol extracts were also prepared for chlorinated ethene analysis. Soil samples were expected to have variable organic matter content, which may have resulted in differential partitioning of chlorinated ethenes between samples. Extraction in organic solvent (methanol) was used to enable better recovery of analytes. Methanol extracts were prepared by adding 20 ml of HPLC grade methanol (Sigma Aldrich, St. Louis, MO) to 2 oz (60 ml) septa jars. Frozen soil was added, and the jars were reweighed to determine the soil mass. Aqueous and methanol extracts were gently mixed and stored at 4  C for four days prior to analysis, according to previously described protocol (Olson et al., 2017; Sale et al., 2015). Detection limits for geochemical and contaminants parameters are listed in Table S1. 2.5. Chlorinated ethene analysis Chlorinated ethenes (PCE, TCE, cis-DCE, trans-DCE, 1,1-DCE, and VC) were analyzed on an Agilent 6890 GC with 5937 mass selective

detector using a RTX-VMS column (30 m, 0.32 mm O.D., 1 mm film)(Restek, Bellefonte, PA). Ultra-high purity (UHP) helium (Praxair, Cedar Rapids, IA) was used as the carrier gas at a flow of 1.5 ml/min at a 5:1 split ratio. The oven program was 35  C for 1 min, 30  C/min to 180  C. A standard curve was prepared by dilution of a commercial standard (502/524 Volatile Organics Calibration Mix, Supelco, Bellefonte, PA). Analytical duplicates were analyzed for 10% of samples. d13C Compound specific isotope analysis (CSIA) of the chlorinated ethenes was performed on selected samples from core 1 to confirm that contaminant biodegradation was occurring and that observed concentration decreases were not simply the result of physical processes (e.g., dilution, sorption). Approximately 5 g of frozen soil was added to 40 ml vials with deionized water (no headspace). The sample was allowed to equilibrate overnight, and the aqueous phase was analyzed for PCE, TCE, cis-DCE, and VC by Pace Analytical (Pittsburg, PA) using a gas chromatograph-isotoperatio mass spectrophotometer. d13C is reported in per mil, (‰) relative to Pee Dee Belemnite.

2.6. Dissolved gas analyses UHP nitrogen (5 ml) was injected into the aqueous extract jar by syringe, then a 5 ml water sample was withdrawn and transferred to a 22 ml nitrogen-purged headspace vial (Restek, Bellefonte, PA). Vials were briefly equilibrated to room temperature (20e22  C). Oxygen was analyzed on an Agilent 6890 GC with thermal conductivity detector and a Supelco 5A molecular sieve column (5 ft x 1/8 inch, 60/80 mesh). Samples (200 mL) were collected with a gastight syringe and manually injected. UHP helium was used as the carrier gas. Separation was isothermal at 70  C. Analytical duplicates were analyzed for 10% of samples. Reduced gases (i.e. methane, ethene, ethane, and acetylene) were analyzed on an Agilent 6890 GC with a flame-ionization

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detector and a 1% SP-1000 on Carbopack B column (6 ft x 1/8 in ID, 60/80 mesh). Separation was isothermal at 90  C. Analytical duplicates were analyzed for 10% of samples. 2.7. Geochemical analysis Geochemical analysis was performed on the aqueous soil extracts. Chloride, nitrate, and sulfate were analyzed by EPA method 300.1 using a Dionex 2100 ion chromatograph with an AS-18 column and isocratic separation (36 mM KOH eluent). Total and ferrous iron samples were filtered (0.45 mm, Millipore, Burlington, MA) and measured spectrophotometrically by Hach methods 8008 and 8146. pH was measured by EPA method 100.1 using a ThermoOrion 9107BN probe. Total organic carbon (TOC) was measured in acidified (pH < 2) aqueous extracts by SM 5310b using a Shimadzu TOC-V analyzer. Total solids were determined gravimetrically by drying soil from the aqueous extracts at 105  C overnight. Total solids were used to normalize analyte concentration in aqueous or methanol phases to the soil dry weight. Analytical duplicates were analyzed for 10% of samples. 2.8. Nucleic acid extraction, real-time PCR (qPCR), and reversetranscription (RT)-qPCR

checked for qPCR product size by agarose gel electrophoresis. In addition, to verify the specificity of qPCR primer sets RTC and RTE, clone libraries were constructed from the amplification products of DNA extracted from core 1 (section 3-2 (2.2m) and 6e8 (4.6m)). Briefly, the qPCR products (with RTC and RTE primer sets) were subjected to a second amplification using Taq PCR Master Mix Kit (Qiagen, Germantown, MD) with thermocycler conditions same as qPCR and additional 10 min elongation step at 72  C. Purified PCR products from the second amplification were ligated into the pCR 2.1-TOPO vector using the TOPO TA Cloning Kit (Invitrogen Corp., Carlsbad, CA) and transformed into One Shot TOP10 chemically competent E.coli cells (Invitrogen Corp., Carlsbad, CA). Recombinant E.coli were plated on Luria Broth agar containing kanamycin (50 mg/L) and X-gal (0.4 mg/plate) and incubated overnight at 37  C (Liang et al., 2015). Clones with successful insertions were Sanger sequenced at the Iowa Institute of Human Genetics Genomics Division with M13 F/R primers (Table S2). For RTC primer set, 4 unique sequences were obtained from 5 clones and were 94e97% similar to etnC from Mycobacterium by BLAST. For RTE primer set, 13 unique sequences were obtained from 13 clones and were 89e99% similar to etnE from Mycobacterium, Nocardioides, and enrichment cultures clones by BLAST (Table S4). 2.10. Statistical analyses

For groundwater samples, DNA and RNA were extracted from Sterivex-GP filters (in duplicates) using the MoBio PowerWater Sterivex DNA Isolation Kit and Mobio PowerWater RNA Isolation Kit, respectively (Mobio, Carlsbad, CA) as described previously (Liang et al., 2017b). Control luciferase mRNA (1 ng) (GenBank accession No. X65316, Promega, Madison, WI) was added to RNA samples after the cell lysis step. After extraction, RNA samples were subjected to contaminating DNA removal, purification, and reverse transcription to cDNA as previously described (Liang et al., 2017b). For cryo-core soil samples, duplicate DNA extractions were performed (approximately 0.25 g each) with the DNeasy PowerSoil Kit (Qiagen, Germantown, MD). DNA extract concentrations ranged from below detection (<0.05 ng/ml) to 21.4 ng/ml, with an average of 1.37 ng/ml as measured by the Qubit dsDNA HS Assay Kit (Thermo Fisher Scientific, Waltham, MA). DNA and cDNA extracts were stored at
80  C prior to further analysis. The abundances of total bacterial 16S rRNA genes and specific functional genes etnC, etnE, mmoX, pmoA, bvcA, and vcrA were estimated as described previously (Liang et al., 2017b). Primer sets for PCR and qPCR are provided in Table S2. Each 20 mL qPCR reaction contained 10 mL Power SYBR Green PCR Master Mix (Invitrogen, Carlsbad, CA), variable primer and template concentrations, and bovine serum albumin (100 ng/ml) (Table S3). All qPCRs were performed with an ABI QuantStudio 7 Flex Real-Time PCR System (Applied Biosystems, Grand Island, NY). PCR thermocycler conditions were: 10 min at 95  C, followed by 40 cycles at 95  C (15 s) and 60  C (1 min) followed by a PCR product dissociation step. Additional qPCR information, including primer and template concentrations, qPCR linear range, standard curve efficiencies, and Y-intercepts are listed in Table S3, in accordance with Minimum Information for Publication of Quantitative Real-Time PCR Experiments (MIQE) guidelines (Bustin et al., 2009). Gene abundances masured by qPCR were normalized to soil dry weight. 2.9. qPCR quality assurance The specificity of SYBR Green based qPCR was validated by dissociation curve analysis as previously described (Liang et al., 2017b). In no template controls, the target gene was either not detected or was detected as primer dimers. Selected samples showing melting curve peaks different from standards were

All statistical analyses were performed in RStudio Version 1.0.153 (RStudio Team, 2015). The relationships between gene abundances and geochemical parameters were assessed by nonparametric Spearman's correlation, simple linear regression (slr), and censored regression (cr). For regression analyses, both gene abundances and geochemical parameters were log transformed to achieve better normality (Figs. S1eS3). To facilitate logtransformation, non-detects in geochemical parameters were replaced by 0.5  method detection limit and corrected for solids content. Censored regression was employed because gene abundance (except for 16S rRNA gene) datasets contained a high percentage of non-detects (26e67%) (Fig. S1) and were thus considered censored datasets. Akaike's information criteria (AIC) scores were calculated for quantitative comparison between simple linear regression and censored regression. Censored regression was performed with the function censReg() (Henningsen, 2017) in the censReg package in R. Overall, data structure and AIC scores were taken into consideration when choosing the most appropriate model for statistical analyses. The non-parametric Wilcoxon signed-rank test was used to make comparisons among different gene abundances. 3. Results 3.1. Stratigraphy Visual observation of the soil samples shows distinct soil layers that are horizontally continuous. Each core consists primarily of an upper silty layer extending from the top of the cores to approximately 3.4e3.7 m, and a lower sandy layer extending to 5.7 m bgs (Fig. S4). These layers were previously characterized as the upper and lower surficial aquifers (Churchill, 2012; Vroblesky et al., 2009), although there is no low-permeability zone separating these layers. The bottom of the aquifer is bounded by layers of peat and clay of variable thickness, which was previously identified as a confining unit (Churchill, 2012). 3.2. Distribution of chlorinated ethenes and ethene with depth The distribution of chlorinated ethenes and ethene in core

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samples was analyzed to identify regions of the aquifer where VC may be generated. VC concentrations in core 1, in the source area, steadily increased from 0.66 to 2.2 m bgs, forming a large bifurcated peak between 2.2 and 2.6 m bgs, then rapidly dropped off by 3.1 m bgs (Fig. 2). A second peak in VC occurred between 5.5 and 5.8 m bgs, immediately above the confining layer. VC concentrations in the other cores, which are located down gradient of the source area, were much lower (<1 mg/kg) than in core 1. Cores 2e3 each had a single peak in VC between approximately 2.8e3.7 m bgs. Core 4 also had a single peak, (3.7e4.6 m bgs). Each of the regions with elevated VC concentrations was found near areas with elevated cisDCE (Fig. 2). In cores 1e3, the spikes in VC were concurrent with spikes in cis-DCE but had a narrower vertical distribution. In core 4, the spike in VC was flanked by regions of elevated cis-DCE but coincided with a steep drop in cis-DCE. Concentrations of trans-DCE and 1,1-DCE were below detection in all samples. The region with the highest VC (Fig. 2) occurred just below the location with the highest PCE and TCE concentrations (25,000 and  530 mg/kg, respectively)(Fig. 2). Spikes in PCE and TCE near the bottom of cores 2e4 did not have corresponding spikes in VC. In cores 1 and 3, peaks in ethene were also coincident with spikes in VC (Fig. 2). Core 3 had an additional region of elevated ethene between 0.66 and 2.1 m bgs in the absence of VC. Cores 2 and 4 had little or no ethene. CSIA was only performed on samples from core 1 to assess the possibility that chlorinated ethenes had been subject to

285

degradation processes. The measured d13C values for PCE were
28‰ (Table 1), within the range of newly manufactured PCE (Hunkeler et al., 2009), while d13C values for both TCE and cis-DCE were higher than for PCE. The samples collected at 0.94 m had a VC d13C of
24‰, higher than for cis-DCE (
27‰) in the same sample. The increasing d13C values measured for the lower chlorinated ethenes (VC and cis-DCE) is indicative of degradation processes (e.g. biodegradation and/or abiotic reactions catalyzed by ZVI).

3.3. Terminal electron acceptors with depth Dissolved oxygen (DO) and nitrate concentrations were below detection (see detection limits Table S1) in all samples (data not shown). In core 1, sulfate was not detected in samples between 0.66 and 1.0 m bgs and remained near zero throughout the upper surficial aquifer (Fig. 3). Sulfate concentrations then rebound in the Table 1 Core 1 chlorinated ethene stable isotope fractionation ratios (d13C, ‰) for VC, cisDCE, TCE and PCE. ND: not detected. Sample depth (m bgs)

VC

cis-DCE

TCE

PCE

0.94 4.0 5.4


24
32
24


27
25
22


28
25 ND


28
28 ND

Fig. 2. Chlorinated ethene depth profiles for each aquifer sediment cryo-core. Top panels: VC (red circles) and ethene (blue squares). Bottom panels: PCE (green circles), TCE (blue squares), and cis-DCE (orange triangles). The approximate depth of the water table (solid line) and the boundary between the upper and lower surficial aquifers (dashed line) are indicated. Concentrations are normalized to soil dry weight. Error bars show range of analytical duplicates (10% of total samples). Error bars smaller than their symbols were omitted. (For interpretation of the references to colour in this figure legend, the reader is referred to the Web version of this article.)

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lower surficial aquifer, steadily increasing between 3.7 and 4.9 m bgs. Ferrous iron steadily increases between 0.97 and 2.2 m bgs, then briefly spikes to 73 mg/kg between 2.5 and 2.8 m bgs before decreasing sharply at 3.1 m bgs (Fig. 3). Ferrous iron concentrations remain relatively low (5 mg/kg) for the remainder of the lower surficial aquifer, except for a small spike (8.6 mg/kg) at 5.5 m bgs. Methane concentrations were low throughout core 1, but were generally higher in the lower surficial aquifer (8.5 mg/kg) than in the upper surficial aquifer (2.6 mg/kg)(Fig. 3). In core 2, sulfate was detected in all samples, but the concentration had a minimum of <5 mg/kg between 1.1 and 1.3 m bgs. In core 2, ferrous iron peaked between 1.6 and 2.8 m bgs, with a maximum of 17 mg/kg. Sulfate was detected throughout cores 3 and 4, Ferrous iron was <1 mg/L throughout cores 3 and 4. 3.4. Total organic carbon (TOC), chloride, and pH values with depth in cores Core 1 had very high TOC concentrations (55e410 mg/kg) which trended higher with depth (Fig. S5). This is likely a residual from a pilot study in which emulsified zero valent iron (ZVI), which contained emulsified vegetable oil, was injected nearby (Su et al., 2012). TOC concentrations in cores 2e4 were generally lower than in core 1. The very high TOC concentration (>30,000 mg/kg) seen at 4.9 m bgs in core 4 is likely a contribution of the peat observed near the confining layer (Fig. S4). Chloride concentrations (Fig. S5) followed trends similar to the chlorinated ethenes (Fig. 2). The highest concentration of chloride (300 mg/kg) was observed between 0.66 and 3.1 m bgs in core 1. As with the chlorinated ethenes, a second peak in chloride was seen between 4.9 and 5.5 m bgs. pH values generally varied between 5 and 7 SU, with no apparent trends between cores, or with depth (Fig. S6). A small number of samples (six out of 28 samples from core 1), mostly between 4.0 and 4.9 m bgs in core 1had an elevated pH (8.7e11 SU). Ethane and acetylene were below detection in all samples (<15 ppmv). 3.5. Abundance of total 16S rRNA genes, and etheneotroph, methanotroph and anaerobic VC-dechlorinator functional genes in core samples Total 16S rRNA gene abundances and depth profiles were similar

among the four cores, with an average of 4.0  107 copies/g soil, ranging from 1.3  102e1.1  109 gene copies/g soil (Fig. S7). Bacteria were most abundant near the ground surface, decreased with depth, and appeared to stabilize at 106 gene copies/g soil below 3 m (Fig. S7). Etheneotroph functional genes were detected in 74% (etnC) and 67% (etnE) of the samples, averaging 1.1  107 and 1.5  106 copies/g soil, respectively (Fig. 4). In all cores, etnC and etnE abundance decreased with depth to 1.3e1.9 m bgs, then either remained steady, or rebounded in the lower surficial aquifer (>3.1 m bgs)(Fig. 4). Methanotroph functional genes were detected in 67% (mmoX) and 46% (pmoA) of the soil samples. Methanotroph and etheneotroph functional gene abundance were positively correlated (Spearman's correlation, p < 0.05). The abundance of mmoX and pmoA was less than etnC and etnE in all cores (Wilcoxon signed rank test, p < 0.05). The highest methanotroph functional gene abundances were found near the ground surface and decreased with depth to 1.3e1.9 m bgs before rebounding in the lower surficial aquifer (Fig. 4). VC reductive dehalogenase genes were detected in 53% (bvcA) and 33% (vcrA) of the soil samples, with averaging 1.2  105 and 1.2  104 copies/g soil, respectively. These reductive dehalogenase genes were detected in 59% of the soil samples. VC reductive dehalogenase functional genes were generally less abundant than those for etheneotrophs (Wilcoxon signed rank test, p < 0.05). Among the two reductive VC dehalogenase genes, bvcA was more abundant overall than vcrA (Wilcoxon signed rank test, p < 0.05), although vcrA and bvcA abundance were similar in core 1 (4.6e5.0 m) and core 4 (3.7e5.4 m) Coexistence of functional genes associated with aerobic and anaerobic VC biodegradation was commonly observed in the core samples. Etheneotroph functional genes (etnC and/or etnE), methanotroph functional genes (mmoX and/or pmoA), and VC reductive dehalogenase genes (bvcA and/or vcrA) coexisted in 48% of the samples, while etheneotroph functional genes and VC reductive dehalogenase genes coexisted in 55% of the samples. The coexistence of etheneotroph and VC reductive dehalogenase genes was also confirmed by their positive correlation (Spearman's correlation, p < 0.001) (Fig. S8). RNA extraction was not successful for the aquifer sediment samples (5 g). However, groundwater samples collected from

Fig. 3. Terminal electron accepting condition depth profiles for each soil core. Iron-reducing conditions are indicated by Fe2þ concentrations (orange circles), sulfate-reducing conditions are indicated by the lack of SO2
4 (pink squares), and methanogenic conditions are indicated by methane concentrations (blue triangles). The approximate depth of the water table (solid line) and the boundary between the upper and lower surficial aquifers (dashed line) are indicated. Concentrations are normalized to soil dry weight. Error bars show range of analytical duplicates (10% of total samples). Error bars smaller than their symbols were omitted. (For interpretation of the references to colour in this figure legend, the reader is referred to the Web version of this article.)

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Fig. 4. Functional gene depth profiles for each cryo-core. Top panels: Etheneotroph (etnC, etnE), Middle panels: methanotroph (mmoX, pmoA) and Bottom panels: VC reductive dehalogenase genes (bvcA, vcrA). The approximate depth of the water table (solid line) and the boundary between the upper and lower surficial aquifers (dashed line) are indicated. Gene abundances are normalized to soil dry weight. Error bars show range of duplicate measurements.

monitoring wells within the contaminant plume contained abundant etheneotroph and VC-dechlorinator functional gene transcripts (108 copies/L) (Fig. S9). Transcripts of bvcA were relatively more abundant in the wells near the source area (ML-2 and PMW4), while vcrA, etnC, and etnE were present at a similar abundance in ML-2, and vcrA is much less abundant in PMW-4. The upper surficial aquifer, in the well farthest from the source area (ML-07-SU), had the lowest concentration of all the functional gene transcripts, with etnC, etnE, and vcrA all below detection. The lower surficial aquifer well in this location (ML-07-SL) contained both etheneotroph and VC-dechlorinator functional gene transcripts at concentrations similar to the up-gradient wells.

3.6. Relationships between functional genes and site geochemical parameters in cryo-core samples We used different statistical methods (censored regression, simple linear regression, and Spearman's correlation) to assess the relationships between functional genes and geochemical parameters. Censored regression was performed to handle the frequent non-detect values present in the functional gene abundance datasets. Simple linear regression and Spearman's correlation was used for comparison with censored regression. The results of all three methods were generally similar, yet differences were noted in some cases (Tables S5eS11). Statistical analyses were performed

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separately on data from the upper and lower surficial aquifers, considering their different geology and hydrogeology (Vroblesky et al., 2009)(Fig. S4). Etheneotroph functional genes (etnC and etnE) were positively correlated with VC concentrations in the lower surficial aquifer (cr, p < 0.05) but not in the upper surficial aquifer (Table 2). VC reductive dehalogenase functional genes (bvcA and vcrA) were strongly related to VC in the lower surficial aquifer (cr, p < 0.001), but only bvcA was related to VC in the upper surficial aquifer (cr, p ¼ 0.002). No significant association was observed between methanotroph functional genes (mmoX and pmoA) and VC concentrations in the upper or lower surficial aquifer. Total bacterial 16S rRNA genes was positively associated with VC concentrations with depth in both the upper and lower surficial aquifer samples (slr, p ¼ 0.004). We also investigated the relationships between concentrations of other chlorinated ethenes (cis-DCE, TCE, and PCE) and VC biodegradation functional gene abundance. Etheneotroph functional genes were positively associated with cis-DCE in the lower surficial aquifer samples (Table S6), and were positively associated with TCE (Table S7) and PCE (Table S8) in the upper and lower surficial aquifers, except for etnE and TCE. For VC reductive dehalogenase genes, bvcA was positively associated with cis-DCE (Table S6) and PCE (Table S8) along depth of both aquifers, while vcrA was positively associated with cis-DCE in lower surficial aquifer (Table S6). For methanotrophs, positive associations were observed for mmoX with cis-DCE in lower surficial aquifer (Table S6), and with TCE and PCE in both aquifers (Tables S7 and S8). PCE, TCE, and cis-DCE were positively associated with total bacterial 16S rRNA genes in both upper and lower surficial aquifers, while cis-DCE was related to total bacterial 16S rRNA gene only in the lower surficial aquifer. Ethene concentrations were positively associated with etnC, etnE, mmoX, bvcA, and vcrA in the lower surficial aquifer (Table S9). Ethene was also positively correlated with bvcA in the upper surficial aquifer (Table S9). Methane concentrations were negatively associated with etnC, etnE, mmoX, pmoA and vcrA in the upper surficial aquifer (Table S10). Methane was positively associated with bvcA through both aquifers and with vcrA in the lower surficial aquifer (Table S10). TOC was positively correlated to total bacterial 16S rRNA genes in both upper and lower surficial aquifers (Table S11). In upper surficial aquifer, TOC was positively associated with etnC, mmoX and vcrA. In lower surficial aquifer, TOC was positively correlated

with etnC, etnE, mmoX, bvcA, and vcrA. 4. Discussion 4.1. Geochemical conditions and evidence of biodegradation Cryo-coring was successful in capturing data trends that reveal steep biogeochemical gradients, and pockets of microbial activity that would not be evident from typical monitoring well groundwater samples. Our data show that this site has complicated and variable geochemical environments, which are likely governed by the site hydrogeology. High spatial resolution trends in geochemistry and biomarkers were evident in each of the cores, but the connections between the cores were more difficult to interpret. The highest PCE and TCE contamination was at the upper surficial aquifer of core 1, which was the closest to the source zone (Fig. 2). Farther along the soil core transect, PCE concentrations decreased, while TCE, cis-DCE, and VC accumulated in further and deeper parts of the aquifer (Fig. 2). These trends could suggest chlorinated ethene dechlorination was occurring along the movement of the plumes, but could also indicate that the sampling transect was not along the centerline of the chlorinated ethene plume. Comparison with historical data (Vroblesky et al., 2009) suggests our sampling transect was parallel to the plume, but was possibly off centerline as evidenced by the difference in chloride values in core 3 samples as compared to the other cores. Groundwater contours also curve near the end of our transect as a result of groundwater entering a collapsed storm drain, further complicating our understanding of plume movement in the groundwater (Vroblesky et al., 2009). Vertical concentration profiles in the cores showed clear evidence of biodegradation in the vicinity of the source area. The large bifurcated VC peak between 2.2 and 3.0 m observed in core 1, which coincided with an ethene peak provides further evidence of complete dechlorination of PCE to ethene at this site. This is despite relatively low pH across most of the site, particularly in core 1 between 0.97 and 3.4 m bgs, where the pH ranges between 4.4 and 5.9 SU (Fig. S6). This is well below the pH range for dechlorination €ffler et al., 2013). In by known Dehalococcoides sp. (pH 6e8 SU)(Lo addition, CSIA data from core 1 showed PCE had more negative d13C values than for TCE, cis-DCE, and VC, providing a second line of evidence that anaerobic reductive dechlorination was occurring (Table 1). Terminal electron acceptor concentrations showed clear differences in the redox conditions between the upper and lower regions

Table 2 Regression analysis of gene abundances and VC in upper and lower surficial aquifer. Abbreviations: Coef: regression coefficient (depicts either positive or negative correlation), NA: not applicable; p values < 0.05 are in displayed in bold. upper surficial aquifer, n ¼ 68

16S rRNA gene etnC etnE mmoX pmoA bvcA vcrA

VC

cis-DCE

TCE

PCE

Coef

p value

95% CI

Coef

p value

95% CI

Coef

p value

95% CI

Coef

p value

95% CI

NA 0.41 0.16 0.17
0.03 1.05 0.00

NA 0.236 0.578 0.345 0.936 0.002 0.989

NA
0.27-1.10
0.41-0.73
0.19-0.54
0.83-0.77 0.40e1.70
0.43-0.42

NA 0.28 0.06 0.09
0.14 0.64
0.03

NA 0.368 0.829 0.604 0.699 0.043 0.857

NA
0.33-0.89
0.45-0.56
0.24-0.41
0.84-0.57 0.02e1.25
0.41-0.34

NA 1.00 0.52 0.50 0.73 0.72 0.24

NA 0.001 0.048 0.003 0.052 0.029 0.224

NA 0.40e1.60 0.00e1.04 0.17e0.83
0.01-1.47 0.07e1.37
0.14-0.62

NA 0.99 0.58 0.46 0.67 0.55 0.22

NA <0.001 0.023 0.005 0.063 0.083 0.247

NA 0.41e1.58 0.08e1.09 0.13e0.78
0.04-1.38
0.07-1.17
0.15-0.59

lower surficial aquifer, n ¼ 49

Coef

p value

95% CI

Coef

p value

95% CI

Coef

p value

95% CI

Coef

p value

95% CI

16S rRNA gene etnC etnE mmoX pmoA bvcA vcrA

NA 0.33 0.58 0.14 0.03 1.22 1.83

NA 0.046 0.024 0.151 0.915 <0.001 <0.001

NA 0.01e0.65 0.08e1.09
0.05-0.33
0.53-0.59 0.76e1.69 0.97e2.70

NA 0.62 1.09 0.33 0.29 1.38 1.55

NA 0.005 0.002 0.011 0.461 <0.001 0.015

NA 0.18e1.05 0.41e1.77 0.08e0.59
0.49-1.08 0.65e2.10 0.30e2.79

NA 0.44 0.59 0.33 0.99
0.59
0.83

NA 0.019 0.066 0.006 0.009 0.083 0.098

NA 0.07e0.80
0.04-1.22 0.10e0.56 0.25e1.72
1.26-0.08
1.82-0.15

NA 0.58 0.56 0.29 0.39 0.75 0.72

NA <0.001 0.036 0.003 0.182 0.010 0.116

NA 0.28e0.88 0.04e1.09 0.10e0.48
0.18-0.96 0.18e1.31
0.18-1.61

P.M. Richards et al. / Water Research 157 (2019) 281e291

of the surficial aquifer. Sulfate and iron reduction were evident in the upper surficial aquifer in cores 1 and 2. The lower surficial aquifer did not show signs of iron or sulfate reduction. This suggests that the groundwater in the lower surficial aquifer did not transit the upper surficial aquifer but rather recharged from a distant location. Although no impermeable layer separates the upper and lower surficial aquifer, there were large differences in observed chemical and hydrological characteristics. Neither oxygen nor nitrate were detected in any samples, indicating oxygen-limiting conditions throughout the aquifer. 4.2. Abundance and distribution of biomarkers Geochemical and CSIA data alone in the core samples could not differentiate between aerobic and anaerobic VC biodegradation processes. However, the potential for in situ aerobic VC degradation in soil samples from Site 45 is demonstrated by estimating the abundance of related functional genes and transcripts (i.e. etnC, etnE, mmoX, and pmoA) using qPCR techniques. We also estimated the abundance of anaerobic VC reductive dehalogenases. This is the first field study to investigate the distribution of three VCdegrading bacterial guilds at high spatial resolution through the use of cryogenic soil coring. Etheneotrophs and anaerobic VC dechlorinators were present throughout the site, and were found to coexist in 55% of soil samples. Specifically, functional genes from etheneotrophs and VCdechlorinators were at similar abundance (103-105 copies/g soil) in samples from core 1 (2.2e2.5 m bgs, 3.7e4.9 m bgs) and core 4 (3.7e5.4 m bgs), which were predominantly anaerobic. The overlapping distributions of aerobic etheneotrophs and anaerobic VCdechlorinators suggests the presence of aerobic and anaerobic microenvironments, or temporal changes in aquifer geochemistry that allow these microorganisms to coexist in very close spatial proximity (within 0.25 g of unhomogenized soil). The high spatial resolution of our soil sample analysis confirmed previous reports of the co-existence and expression of functional genes from aerobic and anaerobic VC-degrading bacteria in anoxic groundwater samples (Liang et al. 2017a, 2017b), and demonstrates that their simultaneous detection is not an artifact of groundwater sampling techniques. This is an important finding, as previous reports of the overlapping presence of etheneotrophs and anaerobic VCdechlorinator functional genes in situ relied on groundwater sampling (Liang et al. 2017a, 2017b). The use of cryogenic coring is the best possible way to prevent fluids distribution during sample recovery, and as a consequence our data represent the most compelling demonstration to date that aerobic and anaerobic VCdegradation bacterial groups coexist in small sediment samples. The presence of functional genes from etheneotrophs and anaerobic VC-dechlorinators in anaerobic portions of the aquifer does not necessarily indicate that these microorganisms were active. Unfortunately, RNA extraction from soil samples (5 g) was unsuccessful, likely as a result of low biomass in the samples (average DNA concentrations 209 ng/g soil) and the instability of RNA during extraction. However, etheneotroph and VCdechlorinator functional gene transcripts were detected in nearby contemporaneous groundwater samples by virtue of much larger sample size (3 L groundwater). Taken together, the presence of etheneotroph functional genes and VC reductive dehalogenase genes in soil and transcripts of these genes in adjacent groundwater suggest that both aerobic and anaerobic VC-degrading bacteria were both present and active throughout the aquifer. 4.3. Correlations between biomarkers and geochemical conditions Etheneotroph functional genes were only correlated to VC

289

concentrations in the lower surficial aquifer. Etheneotrophs can oxidize VC via co-metabolism with ethene as the primary growth substrate (Freedman and Herz, 1996), and some etheneotrophs use VC as sole carbon source under laboratory conditions (Jin and Mattes, 2008). VC and ethene concentrations both showed positive relationships with etnC and etnE abundances in the lower surficial aquifer. This further supports similar relationships noted in groundwater samples (Liang et al., 2017b) and suggests that both VC and ethene supported the growth of etheneotrophs in portions of the lower surficial aquifer. The absence of associations of etheneotroph functional genes with VC in the upper surficial aquifer could be explained by etheneotrophs being more strongly influenced by other factors, such as DO levels, which decreased with depth from the unsaturated zone (Lee et al., 2015; Noll et al., 2005). Although oxygen in all soil samples were below detection limit (0.13 mg/L), sulfate was abundant at the top of the cores, and declined with depth, suggesting that a more favorable terminal electron acceptor was available in shallow portions of the core, possibly a low-level flux oxygen that went undetected. In the lower surficial aquifer, vertical oxygen flux would not be expected, and terminal electron accepting conditions may be relatively constant, allowing the influence of VC concentrations on etheneotrophs to be revealed. The positive relationship between VC concentrations and VC reductive dehalogenase gene abundances in the lower surficial aquifer could be explained by anaerobic VC dechlorinators using VC as electron acceptor (He et al., 2003). Among the two VC reductive dehalogenase genes, bvcA was generally more abundant than vcrA. This observation is consistent with previous studies where both genes were tracked in microcosms and at a contaminated site (Atashgahi et al., 2013; Lee et al., 2008). Only bvcA was correlated to VC concentrations in the upper surficial aquifer samples. The lack of correlation between vcrA and VC in the upper surficial aquifer could be explained in part to pH in this region of the site being well below € ffler et al., 2013). A the optimum for Dehalococcoides (6e8 SU) (Lo previous study which suggests that anaerobic VC dechlorinators employ bvcA for growth on VC in less reducing conditions (van der Zaan et al., 2010). However, high TOC concentrations, apparent sulfate depletion with depth, and the presence of ferrous iron indicate relatively strong reducing conditions in the upper surficial aquifer samples. The positive association between TOC and total bacteria and VC-degrading bacteria indicate TOC (most likely derived from previous emulsified ZVI injections), supported overall bacterial growth at Site 45. VC-oxidizing bacteria and their functional genes have previously been observed in oxygen-limited subsurface environments and in microcosms containing material collected from such environments (Fullerton et al., 2014; Gossett, 2010; Liang et al., 2017b). Extensive searches have failed to reveal any organism capable of VC oxidation with an electron acceptor other than oxygen (Freedman et al., 2013). Several VC-oxidizing isolates display very low oxygen half-velocity constants (0.07e0.3 mg/L)(Coleman et al., 2002), and thus the ability to scavenge oxygen under very low DO conditions. It has been suggested that the systems where anaerobic VC oxidation was reported were not truly anaerobic, but experienced a cryptic low-level oxygen flux (Freedman et al., 2013; Gossett, 2010). Indeed, sustained VC oxidation at DO levels below 0.02 mg/L have has been demonstrated in microcosms provided with a low-level DO flux (Gossett, 2010). This suggests that VC-oxidizing bacteria can exist at the interface of strongly anaerobic subsurface regions. The significant heterogeneity observed at some contaminated sites suggest that very low DO zones are intermingled with truly anaerobic zones, allowing for very close proximity of aerobic and anaerobic VC-degrading bacteria. The possibility that low level DO fluxes can occur at depth in contaminated aquifers and sustain

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P.M. Richards et al. / Water Research 157 (2019) 281e291

aerobic bacterial populations warrants further study. In contrast to functional genes from etheneotrophs and anaerobic VC dechlorinators, methanotroph functional genes were not correlated with VC concentrations in either the upper or lower surficial aquifer samples. This is consistent with previously reported groundwater data (Liang et al., 2017b), and we suspect this is related to the fact that methanotrophs only co-metabolize VC. Overall, methanotroph functional genes were of relatively low abundance at Site 45 (as compared to a previous study of groundwater samples from a dilute VC plume (Liang et al., 2017a)), suggesting that VC co-metabolism by methanotrophs was not a major pathway for VC attenuation. The negative correlations between methane and methanotroph functional genes in the upper surficial aquifer could be explained by aerobic growth of methanotrophs at the expense of methane. The absence of correlations of mmoX, and the weak correlations of pmoA with methane in the lower surficial aquifer suggest that geochemical conditions in the lower surficial aquifer were not ideal for growth of methanotrophs. Associations between cis-DCE concentrations and VC biodegradation functional genes were similar to those seen with VC concentrations. These positive associations could be present because cis-DCE is the parent compound of VC and that cis-DCE is used as an electron acceptor by VC dechlorinators carrying bvcA € ffler et al., 2013). The associations between PCE and and/or vcrA (Lo TCE concentrations and VC biodegradation functional genes differed from those seen with cis-DCE and VC concentrations (i.e., they were associated with etheneotroph functional genes but not VC reductive dehalogenases in both the upper and lower surficial aquifers). Similar correlations were observed previously with monitoring well groundwater samples (Liang et al., 2017b). 4.4. Conclusions  Cryogenic coring can capture higher spatial resolution samples than would be possible with monitoring wells, and is effective for capturing high resolution trends in biomarker and geochemical data.  Cryogenic coring eliminates fluids re-distribution during sampling and thus provides the best available evidence for coexistence of functional genes from aerobic and anaerobic VCdegrading bacterial guilds  Functional genes from aerobic VC-degrading bacteria (both etheneotrophs and methanotrophs) are present and widely distributed in an apparently anaerobic aquifer.  Improved methods for isolating RNA from environmental samples are necessary in order to quantify biological activity in the subsurface at a comparable spatial resolution.  An improved understanding of aerobic VC degradation under low DO concentrations and low oxygen flux conditions would be useful in designing and implementing natural attenuation or aerobic bioremediation strategies for site cleanup. Acknowledgements The work was funded by the Environmental Security Technology Certification Program (ESTCP) under project ER-201425. We €ffler for graciously providing the E.coli host thank Dr. Frank Lo strains carrying pCR2.1 TOPO vectors with bvcA and vcrA insertions and Chris Kocur for helpful comments on the manuscript. Appendix A. Supplementary data Supplementary data to this article can be found online at https://doi.org/10.1016/j.watres.2019.03.059.

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