1,2-Dichloroethane Exposure Alters the Population Structure

Article pubs.acs.org/est

1,2-Dichloroethane Exposure Alters the Population Structure, Metabolism, and Kinetics of a Trichloroethene-Dechlorinating Dehalococcoides mccartyi Consortium Koshlan Mayer-Blackwell,† Maeva Fincker,† Olivia Molenda,∥ Benjamin Callahan,§ Holly Sewell,† Susan Holmes,§ Elizabeth A. Edwards,∥,⊥ and Alfred M. Spormann*,†,‡ †

Civil and Environmental Engineering, ‡Chemical Engineering, and §Department of Statistics, Stanford University, Stanford, California 94305, United States ∥ Chemical Engineering & Applied Chemistry, and ⊥Cell and Systems Biology, University of Toronto, Toronto, Ontario M5S 3E5, Canada S Supporting Information *

ABSTRACT: Bioremediation of groundwater contaminated with chlorinated aliphatic hydrocarbons such as perchloroethene and trichloroethene can result in the accumulation of the undesirable intermediate vinyl chloride. Such accumulation can either be due to the absence of specific vinyl chloride respiring Dehalococcoides mccartyi or to the inhibition of such strains by the metabolism of other microorganisms. The fitness of vinyl chloride respiring Dehalococcoides mccartyi subpopulations is particularly uncertain in the presence of chloroethene/chloroethane cocontaminant mixtures, which are commonly found in contaminated groundwater. Therefore, we investigated the structure of Dehalococcoides populations in a continuously fed reactor system under changing chloroethene/ethane influent conditions. We observed that increasing the influent ratio of 1,2-dichloroethane to trichloroethene was associated with ecological selection of a tceA-containing Dehalococcoides population relative to a vcrA-containing Dehalococcoides population. Although both vinyl chloride and 1,2-dichloroethane could be simultaneously transformed to ethene, prolonged exposure to 1,2-dichloroethane diminished the vinyl chloride transforming capacity of the culture. Kinetic tests revealed that dechlorination of 1,2-dichloroethane by the consortium was strongly inhibited by cis-dichloroethene but not vinyl chloride. Native polyacrylamide gel electrophoresis and mass spectrometry revealed that a trichloroethene reductive dehalogenase (TceA) homologue was the most consistently expressed of four detectable reductive dehalogenases during 1,2-dichloroethane exposure, suggesting that it catalyzes the reductive dihaloelimination of 1,2-dichloroethane to ethene.

1. INTRODUCTION Establishing and maintaining a diverse microbial community containing Dehalococcoides spp. is a common strategy for achieving complete biological dehalogenation of chlorinated solvents in contaminated groundwater.1−6 Predicting the relative abundances of organohalide-respiring microorganisms is complicated when multiple classes of organic halogenated cocontaminants are present, the metabolisms of different populations intersect at various biodegradation steps, or intermediates are inhibitory.7−11 Of common concern are mixtures of chlorinated ethenes (e.g., perchloroethene (PCE), trichloroethene (TCE), cisDCE (cis-DCE), and vinyl chloride (VC)) and chlorinated ethanes (e.g., 1,1,1-trichloroethane (1,1,1-TCA), 1,2-dichloroethane (1,2DCA)) that are often copresent in many contaminated sites.12 Because many in situ bioremediation efforts can last years, or even decades, understanding the effects of heterogeneous contaminant mixtures on the long-term population structure and activity of organohalide-respiring bacteria is crucial for groundwater cleanup. © XXXX American Chemical Society

Co-contaminants can influence the degradation of a primary compound by at least two possible mechanisms. The first mechanism is enzymatic, where a cocontaminant may directly inhibit enzymes required to catalyze the transformation of the primary compound.13,14 A second mechanism depends on a shift in available resources, in which the introduction of a cocontaminant results in competition during growth, disadvantaging another microbial subpopulation (or a symbiotic partner).15 For instance, 1,1,1-TCA impacts chloroethene degradation via the enzymatic mechanism by strongly inhibiting the rate of TCE and VC dechlorination in three commercially produced bioremediation cultures,9,11 thereby slowing down metabolic dechlorination of chlorinated ethenes by certain Dehalococcoides mccartyi. Received: June 13, 2016 Revised: October 6, 2016 Accepted: October 12, 2016

A

DOI: 10.1021/acs.est.6b02957 Environ. Sci. Technol. XXXX, XXX, XXX−XXX

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Environmental Science & Technology

500bp. Additionally, an aliquot of the same DNA was sequenced using a nanopore DNA sequencing device (Oxford Nanopore) without prior DNA fragmentation to achieve long reads (1−70kb). We coassembled reads generated by Illumina Hi-Seq (2 × 100 bp) and Oxford Nanopore MinION (1−70 kb) DNA sequencing instruments using the SPAdes assembly algorithm.23 We binned the contigs by mean read depth, a measure of the relative abundance of a given sequence within the community DNA pool. Contigs were annotated for predicted proteins using the Prokka software.24 Contigs assembled using both high depth of coverage Illumina reads with Oxford Nanopore reads as scaffolds were longer than contigs generated from Illumina reads alone (data not shown). 2.3. DNA Extraction. Previously frozen cells, spun down from 4 mL reactor material, were thawed on ice. 250 μL of sterile phosphate buffered solution (MP Biomedicals) was added to the pellet. We resuspended the pellet into solution by mild vortexing for 10 s. Thirty μL detergent solution (MT buffer, MP Biomedicals) was added to the cell slurry to aid cell lysis. We added glass beads (Lysing Matrix E, MP Biomedicals) to each tube. To achieve mechanical lysis, we placed the lysing tubes on a 2 mL tube vortex adapter (MoBio) run at max speed for 10 min at 4 °C. After mechanical lysis, we centrifuged the tubes for 10 min at 16 000 g at 4 °C. We transferred the aqueous phase above the beads to a new 2 mL heavy phase-lock gel tube (5Prime) and applied 250 μL of UltraPure 25:24:1 phenol:chloroform:isoamyl-alcohol (ThermoFisher Scientific). Emulsification of the aqueous and organic phases was achieved by vigorous manual agitation for 20 s before the phase-lock tubes were centrifuged at 16 000 g for 5 min at 4 °C. The aqueous phase was concentrated using a column method according to the manufacturer’s instructions (Clean and Concentrator-5, Zymo Technologies). We subsequently quantified DNA recovery by spectrophotometric means (NanoDrop Thermo Scientific) and by fluorometric assay (Qubit BR dsDNA, ThermoFisher Scientific). A portion of the isolated DNA was then diluted to 10 ng/μL in PCR-grade water and frozen for future analysis. 2.4. Amplicon Library Generation. We generated dualindex barcoded PCR amplicon libraries25 targeting conserved and variable nucleotide sequences in the gene for the large-subunit Ni−Fe uptake hydrogenase (hupL) found in all Dehalococcoides isolates and 16S rRNA gene found in all bacteria. The hupL 257F forward 5′-CGAATAACGGGCGGATACTCC-3′ and hupL 623R reverse 5′-GTAACACCGCCGGCWACC-3′ primers were combined with a sequencing adapter, pad, linker, and unique index sequence (see Supporting Information File 1 for all primer sequences). For the 16S rRNA primers we selected primers amplifying the V1 and V2 regions: 28F 5′-GAGTTTGATCNTGGCTCAG-3′and 338R 5′-TGCTGCCTCCCGTAGGAGT- 3′. These primers have been used successfully in conjunction with a dual-index barcoded approach.26 This region of the 16S rRNA gene was selected for this study, because it contains the most Dehalococcoides strain-specific single nucleotide polymorphisms compared to amplicons targeting the V4, V5, or V6 regions. Despite satisfactory performance of these dualindex primer libraries, future efforts might consider using these hupL specific amplification primers in conjunction with more recently developed Nextera XT DNA Library Prep Index Kit (Illumina) to enhance overall sequencing quality and lower library construction cost. Dual-index amplicon libraries were generated for duplicate DNA extractions from each reactor time-point. Amplicon

However, the role of chloroethanesas resources rather than as inhibitorsin shaping the population structure of Dehalococcoides mccartyi-containing consortia has received less attention. Testing for ecological impacts of noninhibitory chloroethanes requires measuring the frequency of various subpopulations in mixed communities under prolonged exposure to contaminant mixtures at concentrations and growth rates representative of contaminated subsurface environments. In this study we focused on the compound 1,2-dichloroethane (1,2-DCA)a cocontaminant in 634 of 997 chloroethenecontaminated EPA NPL sites in the United States.12 In contrast to 1,1,1-TCA, 1,2-DCA has not been shown to be a strong enzymatic inhibitor of Dehalococcoides mccartyi reductive dehalogenases. However, given that 1,2-DCA can support the growth of a diverse range of organohalide-respiring bacteria,8,16 including some Dehalococcoides mccartyi strains,17,18 we sought to elucidate its ecological effects on a bacterial consortium comprised of multiple chloroethene-respiring populations. The effects of 1,2-DCA exposure on Dehalococcoides dominated microbial consortia are difficult to predict a priori, since studies suggest that multiple reductive dehalogenases found in Dehalococcoides mccartyi strains could mediate the dihaloelimination of 1,2-DCA to ethene. Maymó-Gatell et al. (1999) observed dechlorination activity for 1,2-DCA in Dehalococcoides mccartyi strain 195, although the responsible enzyme was not determined.17 Recently, Tang et al. (2013) demonstrated activity of the Dehalococcoides mccartyi strain BAV1 enzyme BvcA for 1,2-DCA in native polyacrylamide gel electrophoresis (nativePAGE) assays, and Parthasarathy et al. (2015) showed that heterologously expressed Dehalococcoides mccartyi strain VS enzyme VcrA transformed 1,2-DCA to ethene in vitro.19,20 Duhamel and Edwards (2007) previously investigated the response of the bacterial consortium, KB-1, to batch addition of methanol and 1,2-DCA.18 They showed that the ribosomal 16S rRNA gene marker for Dehalococcoides as well as for Geobacter in the KB-1 consortium increased with dihaloelimination of 1,2-DCA to ethene. However, those experiments left uncertain the effects of longer-term 1,2-DCA exposure on the ecological dominance of different Dehalococcoides subpopulations and the vinyl chloride transforming capacity of the culture.

2. MATERIALS AND METHODS 2.1. Reactor Construction. Benchtop reactorsEV-3, EV-TCE, EV-Pwere constructed using 1 and 0.5 L Pyrex bottles modified with a GL45 threaded caps (JR-S-11003, VICI) and polyetherketone (PEEK) fittings. Due to the extreme oxygen sensitivity of Dehalococcoides,21 each reactor was maintained anoxic by positive pressure supplied from a tank of 90:10 N2:CO2 gas and fed fresh medium using a gastight syringe (SGE Analytical) and syringe pump (Harvard Apparatus). Further details of the medium preparation, reactor sampling, and analytical methods can be found in the Supporting Information. 2.2. Metagenomic Sequencing and Assembly. The original EV-5 inoculum was derived from sediment of a chloroethene-contaminated site in Corvallis, OR.22 Immediately prior to inoculating the EV-3 reactor, we isolated total DNA from the parent EV-5 reactor. DNA was sequenced by two technologies. First, we paired-end sequenced (2 × 100bp) total community DNA in a single lane of an Illumina HiSeq sequencing instrument (Genewiz, USA). The sequencing library was prepared using mechanical fragmentation and the TruSeq (Illumina) library preparation method with a target insert size of B


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Environmental Science & Technology generating polymerase chain reactions were comprised of 2.2 μL 10X AccuPrime Pfx Reaction mix, 0.2 μL polymerase AccuPrime Pfx DNA polymerase (ThermoFisher Scientific), 0.2 μL Magnesium Chloride (50 mM), 2 μL forward index primer (10 μM), 2 μL reverse index primer (10 μM), 2 μL DNA template (10 ng/ μL), and 13.4 μL PCR-grade water. The PCR program was as follows: 95 °C for 3:00, and 30 cycles of 95 °C for 2:00, 55 °C for 0:30, 65 °C for 0:30, followed by an extension step at 68 °C for 10:00. PCR reactions were cleaned up using a 96-well column method according to the manufacturer’s instructions (ZR-96 DNA Clean & Concentrator-5, Zymo Research) and quantified with a fluorometric assay (Quant-iT PicoGreen dsDNA Assay Kit, ThermoFisher Scientific) prior to being pooled and paired-end 2 × 250bp sequenced on a MiSeq instrument (Illumina). Raw reads were quality filtered and the perlibrary frequency of each unique allele determined using DADA227 software in the R statistical environment.28 2.5. Native PAGE Gel Analysis. Native polyacrylamide gel electrophoresis (native-PAGE) fractionation, dechlorination activity assays, and mass spectrometry were performed as described previously.19,29,30 Mass spectrometry data was analyzed using Scaffold4 Software (Proteome Software, Inc.), with a database of peptides consisting of all amino acids recovered from EV metagenomic assembly and a curated database of reductive dehalogenases present in NCBI’s database.31 Protein concentration in cell extracts was assayed via a Bradford reagent (BioRad) while protein concentrations in gel bands were estimated using ImageJ program (http://imagej.nih.gov/ij/). Further details can be found in the Supporting Information. 2.6. Batch Kinetic Tests. To assay rates of reductive dechlorination, we performed cell suspension assays. We first prepared concentrated stocks of chlorinated compounds (0.1−10 mM) and sodium formate (20 mM) in sterile anoxic cellular growth medium equilibrated overnight. 60 mL of continuously mixed reactor liquid−containing bacterial cells− was transferred anoxically via PEEK tubing into stoppered serum vials previously flushed with 90:10 N2:CO2. We sparged the cell suspensions with N2:CO2 gas for 10 min to remove residual ethene and chlorinated compounds. For time-series kinetic experiments, we then injected a known volume of the concentrated stock solutions into the batch cell suspension using a sterile N2-flushed syringe. Serum vials were incubated in a fume hood at 21−23 °C. We periodically sampled 1 mL of cell suspension liquid using a N2-flushed syringe and needle. The sampled reactor liquid was immediately injected into a capped 10 mL GC headspace sample vial containing 1.3 mL of 3% hydrochloric acid to halt further enzymatic activity before analysis by gas chromatograph as described in the Supporting Information. For instantaneous rate measurements, we transferred 1 mL aliquots of N2:CO2 − sparged reactor liquid into 10 mL sealed GC headspace vials already containing chlorinated stock solution and formate stock solutions. Work was performed in an anoxic glovebox (3% hydrogen: 97% nitrogen) environment to prevent oxygen exposure during cell transfer. Reaction vials were shaken for 1 h before being halted via the injection of 3% hydrochloric acid solution. 2.7. Kinetic Parameter Estimation. For initial rate studies, we modeled concentration-rate dependence using a Monod model (eq 1) and modified Monod models with either a competitive or noncompetitive inhibition term (eq 2 and 3, respectively), where q is the initial rate, [S] is the initial concentration of substrate, VMAX is the maximum rate of substrate

utilization, KS is the half saturation constant, [I] is the concentration of competitive or noncompetitive inhibitor, and KI is the inhibition constant. ⎛ [S] ⎞ q = Vmax ⎜ ⎟ ⎝ [S] + KS ⎠

(1)

(with competitive inhibition) ⎛ ⎜ [S] q = (Vmax )⎜ ⎜ [S] + KS 1 + ⎝

(

[I ] KI

)

⎞ ⎟ ⎟ ⎟ ⎠

(2)

(with non-competitive inhibition) ⎛ ⎞ ⎜ Vmax ⎟⎛ [S] ⎞ q=⎜ ⎟ ⎜ ⎜ 1 + [I ] ⎟⎟⎝ [S] + KS ⎠ ⎝ KI ⎠

(3)

From initial rate experiments with reactor cell suspensions we estimated VMAX and KS values independently for 1,2-DCA and cis-DCE using the nonlinear Michaelis−Menten regression function SSmicmen() within the R statistical environment.32 We further estimated parameters for substrate inhibition (KI, DCE,DCA and KI, DCA,DCE) using forward simulation of 1,2-DCA batch depletion in the presence and absence of cis-DCE. Depletion of cis-DCE and 1,2-DCA were predicted according to forward simulation of a system of coupled differential equations. Simulation were performed with (i) eq 4 and 5, and (ii) 4 and 6. The simulated change in concentrations were compared to experimental observations of cell suspensions amended with both 1,2-DCA and cis-DCE. (competitive inhibition of cis-DCE by 1,2-DCA) ⎛ ⎜ d[DCE] [DCE] = (Vmax,DCE)⎜ dt ⎜⎜ [DCE] + K S ,DCE 1 + ⎝

(

[DCE] KI ,DCA → DCE

)

⎞ ⎟ ⎟ ⎟⎟ ⎠ (4)

(competitive inhibition of 1,2-DCA by cis-DCE) ⎛ ⎜ d[DCA] [DCA] = (Vmax,DCA )⎜ dt ⎜⎜ [DCA] + K S ,DCA 1 + ⎝

(

[DCE] KI ,DCE → DCA

)

⎞ ⎟ ⎟ ⎟⎟ ⎠ (5)

(non-competitive inhibition of 1,2-DCA by cis-DCE) ⎛ Vmax,DCA d[DCA] ⎜ =⎜ dt ⎜⎜ 1 + [DCE] KI ,DCE → DCA ⎝

(

C

)

⎞ ⎟⎛ ⎞ [DCA] ⎟⎜⎜ ⎟⎟ ⎟⎟⎝ [DCA] + KS ,DCA ⎠ ⎠

(6)


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Environmental Science & Technology Table 1. Reactors in This Studya influent

a

reactor

source

initiation

electron acceptor

EV-5 EV-3 EV-TCE EV-P

Evanite Sediment (OR, USA) EV-5 EV-5 EV-5

1/1/08 7/14/14 5/15/15 5/15/15

TCE (9 mM) TCE (5−0 mM), 1,2-DCA (0−15 mM) TCE (6−9 mM) 1,2-DCA (10 mM)

electron donor Formate Formate Formate Formate

(45 (45 (45 (45

RT

mM) mM) mM) mM)

50 25 25 25

day day day day

RT- Residence Time, TCE- trichloroethene, 1,2-DCA- dichloroethane.

VMAX and KS values were parametrized from the uninhibited initial rates experiments. Using those values, we further calibrated the KI, DCA→DCE and KI, DCE→DCA parameters by minimizing the sum of squared error between modeled and observed concentrations in batch depletion assays, using the Nelder−Mead simplex optimization method33 implemented in the optim() function within the R statistical environment.32

Table 2. Dehalococcides mccartyi Populations in the EV-5 Parent Reactor Metagenome Dehalococcoides mccartyi TceA (DhcEV-TceA) draft genome size contigs read depth (mean) read depth (std. deviation) largest contig number of putative protein coding ORFs number of putative rdhA biochemically characterized rdhA most similar Dehalococcoides mccartyi genome(s) by all vs all blastp rdhA with no similar homologue (>90% amino acid level) in another Dehalococcoides mccartyi genome nif D gene (nitrogen fixation)

3. RESULTS AND DISCUSSION To investigate the ecological effects of 1,2-DCA exposure on Dehalococcoides populations within a chloroethene-dehalogenating consortium, we used a TCE-fed reactor (EV-5), maintained for the 8 years prior, to seed a set of experimental reactors: EV-3, EV-TCE, EV-P (Table 1). A previous study indicated that the parent reactor EV-5 was comprised of multiple populations of organohalide-respiring bacteria,34 and our recent metagenomic analysis revealed that it contained distinct tceAcontaining and vcrA-containing populations. Reactor EV-3 was exposed to a changing mixture of TCE and 1,2-DCA over the course of a year and then fed 1,2-DCA only. The reactor EV-TCE was fed TCE only−analogous to the conditions in the parent reactor EV-5. EV-P was initially fed TCE and then 1,2DCA only for 2 months. By combining molecular observations from a 250-day time course experiment in reactor EV-3 with kinetic tests and enzyme identification by mass spectrometry performed on material from all three reactors, we explored the ecological effects of 1,2-DCA exposure on (i) bacterial community structure, (ii) vinyl chloride transformation potential, and (iii) consortium-level reductive dehalogenase gene expression. 3.1. Initial Dehalococcoides Population Structure. To determine the number of Dehalococcoides or other organohalide-respiring microbial populations comprising the EV-5 inoculum, we sequenced total community DNA from the EV-5 reactor prior to inoculating the EV-3 reactor. Assembled metagenomic contigs, binned by read coverage, generated two distinct near-complete genomes at moderate and high coverages, suggesting that the reactor was dominated by two distinct populations of Dehalococcoides mccartyi (Table 2). We refer to them here as populations DhcEV-TceA and DhcEVVcrA. Population DhcEV-TceA is most closely related to Dehalococcoides mccartyi strains CG4 and 195.35,36 The mean identity of protein coding sequences in DhcEV-TceA based on 1000 homologues with both Dehalococcoides mccartyi strains CG4 and 195 was greater than 96% at the amino acid level. An assembled contig in the DhcEV-TceA bin hosted the gene encoding for a homologue of the trichloroethene reductive dehalogenase gene (tceA), present in Dehalococcoides mccartyi strain 195 but not in strain CG4. A tceA homologue was not detected in DhcEV-VcrA contigs (Table 2). Based on its protein coding sequences, DhcEV-VcrA is most closely related to Dehalococcoides mccartyi strains CBDB1, BAV1, GT, DCMB5, BTF08, and CG5 with mean percent

Dehalococcoides mccartyi VcrA (DhcEV-VcrA)

1.36 Mb 8 2418 284 949 189 bp 1392

1.44 Mb 14 530 117 575 060 bp 1487

13 TceA 195, CG4 2

27 VcrA CBDB1, BAV1,GT,DCMB5, BTF08, CG5 7

yes

no

amino acid identity greater than 96% with homologues in all the above-mentioned strains.36−39 The assembled contigs of population DhcEV-VcrA contained a gene encoding for a previously characterized vinyl chloride reductase (vcrA), not found in the DhcEV-TceA contigs, as well as 26 other putative rdhA genes (Table 2). Despite the high degree of similarity in genomic content to previous isolates, each population assembly revealed a unique set of reductive dehalogenase genes not seen previously in another isolate. Notably, the assembled draft genomes of populations DhcEV-TceA and DhcEV-VcrA contain collectively 9 novel reductive dehalogenase without a highly conserved homologue (>85% at the amino acid level) in any previously available Dehalococcoides mccartyi genome. Based on DNA sequencing depth, Dehalococcoides mccartyi represented greater than 90% of the EV-5 community, at the 16S rRNA gene level. The assembled Dehalococcoides mccartyi 16S rRNA gene sequences had a total read depth of ∼2800, a level of coverage in 20-fold excess to that of any other 16S rRNA gene detected. In other contigs assembled from the metagenome, we detected Desulf itobacterium, Desulfovibrio, Thermovirga, and Spirochete ribosomal sequences present at read depths between 10 and 100. 3.2. Dehalococcoides Population Structure in Response to 1,2-DCA. We tested whether transitioning reactor influent over a 250-day time course from TCE to a mixture of TCE plus 1,2-DCA would impact the consortium’s Dehalococcoides population structure. We analyzed PCR amplicon libraries to estimateat high-temporal resolutionthe relative abundance of Dehalococcoides 16S rRNA and Ni−Fe hydrogenase (hupL) gene variants (Figure 1). The two major Dehalococcoides mccartyi populationsDhcEV-TceA and DhcEV-VcrAexhibited similar D


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Figure 1. Change in Dehalococcoides population structure during increased 1,2-DCA exposure in the EV-3 Reactor. The data represent molecular and chemical measurements over a 250-day reactor time course. Ratios of DhcEV-TceA and DhcEV-VcrA populations were determined by the frequency of (a) 16S rRNA gene variants and (b) of Ni−Fe hydrogenase (hupL) gene variants. Bars represent the range of two measurements generated from two independent DNA extractions and barcoded amplicon pools. (c) Influent concentrations of TCE and/or 1,2-DCA in electron equivalents (1 mM TCE = 6 mM eeq, 1 mM 1,2-DCA = 2 mM eeq). (d) Reactor effluent concentrations of chlorinated ethenes/ethanes. Ethene concentration represents estimated total ethene mass in both the liquid and gas compartments divided by the reactor liquid volume (assuming a Khcc = 7.24). (e) Enlarged view of low effluent concentrations less than 0.5 mM.

fitness, as indicated by their comparable population frequency, during growth on a continuous flux of TCE. Simultaneous 1,2-DCA and TCE influent resulted in continued growth of both-populations; however, the full time course data revealed that the two populations differentially utilized 1,2-DCA. Each increase in the influent 1,2-DCA/TCE ratio resulted in a statistically significant decrease in the relative abundance of the vcrA-host population (Mann Kendall trend test: days 140−165 (p = 0.007), 215−285 (p = 0.02), 285−350 (p = 0.008)). Prolonged growth under a high 1,2-DCA/TCE ratio resulted in ecological dominance of the tceA-host population (Figure 1A and B). During most of that time course, aqueous concentrations of TCE remained below the limit of detection (∼1 μM), and daughter chlorinated ethenes (cis-DCE and VC) were detected at aqueous concentrations less than 10 μM (Figure 1E). By contrast, effluent 1,2-DCA concentrations persisted in excess of 100 μM during most of the time course. This suggests that

the active reductive dehalogenase(s) expressed in the reactor have a higher apparent KS value for 1,2-DCA compared to their respective KS for chloroethenes. The population shift associated with long-term exposure to 1,2-DCA influent was associated with a diminished overall vinyl chloride respiring capacity of the microbial consortia. When determined in batch assays over a concentration range of 25 μM to 300 μM, the VC degradation rate of the EV-TCE reactor derived cells, those never exposed to 1,2-DCA, was 24.6 nmol h−1 per mL cell suspension (95% CI [23.1, 26.2]), whereas the VC dechlorination rate of the EV-3 rector derived cells, those that had experienced 9 months of constant 1,2-DCA exposure, was 3.2 nmol h−1 per mL cell suspension (95% CI [2.1, 4.3]). These findings are consistent with data from an independent study of the EV parent consortium where the observed maximum rate of VC dechlorination ranged over 2 orders of magnitude and was proportional to the relative E


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Environmental Science & Technology abundance of the vcrA-containing population (data not shown). In addition to its negative effect on VC dechlorination rates, we also observed that long-term exposure to 1,2-DCA increased the apparent threshold concentration at which the consortium no longer actively dechlorinated VC (Supporting Figures S1 and S2). 3.3. Dechlorination Kinetics of cis-DCE and 1,2-DCA and Substrate Interference. Substrate interactions were shown previously to impact the rate of stepwise reductive dehalogenation of chlorinated ethenes by Dehalococcoides mccartyi in the mixed microbial consortia.40 Despite the common cooccurrence of 1,2-DCA and chlorinated ethenes, we are not aware of a previous study examining the effect of cis-DCE on the rate at which reductive dehalogenases in a Dehalococcoides sp. catalyze the dehalogenation of 1,2-DCA. To test the hypothesis that cis-DCE inhibits the enzymatic reductive dihaloelimination of 1,2-DCA in this Dehalococcoides-dominated consortium, we performed (i) initial rate experiments (Figure 2), and (ii) monitored the fate of chloroethenes and

Figure 3. Observed and modeled biodegradation of 1,2-DCA by the EV-3 consortium in batch depletion assays. Cell suspensions were amended with (a) 1,2-DCA plus cis-DCE or (b) 1,2-DCA only. Points represent average μmols of each compound per bottle (total mass based on measured aqueous concentrations and the dimensionless Henry’s constants: KH,CC, ETH = 7.24, KH,CC, VC = 1.137, KH,CC, cDCE = 0.167, KH,CC, DCA = 0.0483). Error bars show the standard deviation in triplicate cell-suspensions. The degradation rate of 1,2-DCA was strongly inhibited by the presence of cis-DCE but not VC. Lines represent model results based on coupled Monod differential equations with kinetic parameters from the initial rates experiments (Figure 2). Model scenarios include cases with (i) no substrate inhibition, (ii) dual competitive inhibition (KI,DCE→DCA = 9 μM, KI,DCA→DCE = 45 μM), and (iii) noncompetitive inhibition of DCA dechlorination by DCE (K I,DCE→DCA = 25 μM) and competitive inhibition of DCE dechlorination by DCA (KI,DCA→DCE = 45 μM). KI parameters were calibrated by nonlinear minimization of sum of squared errors between model prediction and experimental observation.

rate of cis-DCE dechlorination (VMAX, DCE) of 112 nmol h−1 mL−1 was greater than VMAX,DCA by nearly a factor of 3, and the KS,DCE value of 8.5 μM was an order of magnitude lower than the value of KS,DCA (Table 3). In addition, when comparing batch 1,2-DCA dehalogenation experiments in the presence and absence of cisDCE, we found that cis-DCE, but not vinyl chloride, strongly inhibited 1,2-DCA dihaloelimination (Figure 3). Including inhibition terms in numerical models (see eq 4-5 in the Materials and Methods) of batch depletion reduced the sum of squared errors by 97% compared with the model not accounting for crosssubstrate inhibition (Table 3, Figure 3). In batch depletion assays, once cis-DCE had been fully converted to vinyl chloride, 1,2-DCA dihaloelimination proceeded at rates similar to the 1,2-DCA only cell suspensions (Figure 3), indicating that the observed 1,2-DCE inhibition was reversible. In contrast, the presence of vinyl chloride in excess of 250 μM aqueous concentration had a negligible effect on the rate of 1,2-DCA dehalogenation (Figure 3). By comparison, the presence of 1,2-DCA only weakly inhibited cis-DCE dechlorination (Figure 3). Additional initial rate experiments, simultaneously varying the starting 1,2-DCA and cis-DCE concentrations, showed strong inhibition of 1,2-DCA dechlorination at initial cis-DCE concentrations of 25 μM and 65 μM (Supporting Figure S3). These results show that

Figure 2. Initial rates of cis-DCE dechlorination and 1,2-DCA dihaloelimination in cell suspensions in the presence of increasing initial substrate concentration. Initial rates of dehalogenation were generated separately for each compound with aliquots of the EV-3 reactor material sampled on the same day of operation. Initial rates were measured in triplicate batch vials. Vertical error bars display the standard deviation about the mean of measured reaction rates. Horizontal error bars display the standard deviation of measured starting aqueous substrate concentration. Dashed lines represent a Monod model fit to each data set based on minimization of sum of squared errors (Table 3).

chloroethanes in batch depletion assays, derived from EV-3, which we either coamended with cis-DCE plus 1,2-DCA (Figure 3a) or amended with 1,2-DCA-only (Figure 3b). All kinetic tests were performed after the reactor EV-3 consortium had been supplied influent containing only 1,2-DCA (absent any chloroethene) for more than 9 months. Using this adapted reactor consortium, initial dechlorination rate measurements using only one substrate, indicated that the EV-3 consortium had preferential dehalogenation activity for cis-DCE versus 1,2-DCA (Table 3, Figure 2). The measured maximum F


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Environmental Science & Technology dihaloelimination of 1,2-DCA by Dehalococcoides mccartyi may not occur in cocontaminated environments, consistent with a similar finding of a recent study by Yu et al. (2013) showing that 1,2-dibromomethane strongly inhibited 1,2-DCA dihaloelimination in another Dehalococcoides mccartyi−containing enrichment culture.41 3.4. Expression of Reductive Dehalogenases During 1,2-DCA Transformation. In order to identify the reductive dehalogenase(s) mediating the dihaloelimination of 1,2-DCA in reactors EV-3 (after 9 months of prior exposure to 1,2-DCA as the sole chlorinated electron acceptor) and reactor EV-P (after 2 months of prior exposure to 1,2-DCA), we used a coupled native polyacrylamide gel electrophoresis (native-PAGE) assay and proteomic approach.19,30 Proteins in the EV-3 and EV-P whole cell extract were separated by molecular mass via native-PAGE and assayed for reductive dehalogenation activity, with separate gel fractions assayed for 1,2-DCA and cis-DCE dechlorination activity, respectively. The highest dechlorination activity for both 1,2-DCA and cis-DCE was found in bands 4 and 5 (Supporting Table S1). Mass spectrometric analysis of these bands identified peptides of four reductive dehalogenases using the translated amino acid sequences from the EV-3 metagenome, as well as a database comprised of publically available reductive dehalogenase protein sequences.31 These peptides were also most abundant in bands 4 and 5. Based on total unique spectrum counts, TceA was the most abundant reductive dehalogenase in both EV-3 and EV-P (Supporting Table S1). In the cis-DCE assays, the major dechlorination product was vinyl chloride. In the 1,2-DCA assays, the major dechlorination product was ethene, consistent with other studies that showed Dehalococcoides sp. transformed 1,2-DCA via dihaloelimination.17,18 Other peptides with a high degree of amino acid identity to (i) VcrA, (ii) a member of Reductive-Dehalogenase-OrthologGroup (RD-OG) 15, and (iii) a member of RD-OG 48 were

also detected in the EV-P derived native-PAGE gel. RD-OG 15 comprises dehalogenases (e.g., DET 1545) that are highly conserved in all Dehalococcoides mccartyi strains known to be expressed during starvation.29,42 RD-OG 48 has not been characterized at this time. Notably, the VcrA, RD-OG 15, and RD-OG 48 peptides were greatly diminished in the EV-3 bands (Supporting Table S1). The diminished presence of VcrA-derived peptide in EV-3 relative to EV-P further supports the hypothesis that persistent exposure to 1,2-DCA gradually selected against the vcrA-containing Dehalococcoides population in this consortium. In direct shotgun proteomics assays from reactor EV-TCE, processed without the native-PAGE gel-fractionation step, peptides from the same four reductive dehalogenase proteins, and no other reductive dehalogenase peptides, were detected (data not shown). This suggests that TceA expression in this reactor system was similar regardless of the presence of TCE or 1,2-DCA in the influent. Although we cannot completely rule out the possibility that distinct dehalogenases catalyze (i) cis-DCE dechlorination to VC and (ii) 1,2-DCA dihaloelimination to ethene, the consistent and numerically dominant signal of TceA peptides in the native-PAGE mass spectrum of EV-3 after 9 months exposure to 1,2-DCA and the observed cross-substrate kinetic inhibition strongly supports the hypothesis that TceA performs both functions. 3.5. Implications for Contaminated Sites. Our current study showed that long-term exposure of the EV-3 consortium to 1,2-DCA favored the growth of a TceA-containing population over a VcrA-containing population. This shift predictably reduced the overall consortium’s maximum rate of vinyl chloride transformation to ethene by an order of magnitude, consistent with the hypothesis that a limited set of vinyl chloride reductases, not present in all strains of Dehalococcoides mccartyi strains efficiently catalyze this final chloroethene degradation step.1,4,6 We infer that increasing the influent ratio of 1,2-DCA to TCE selected for the DhcEV-TceA population over the DhcEV-VcrA

Table 3. Kinetic Parameters from Initial Rate Measurements and Estimates of Inhibition Parameters from Batch Depletion Studiesa

G


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Dehalococcoides as present in the DhcEV-TceA population of the EV consortium are not ideal candidates for the biological removal of 1,2-DCA below the EPA-mandated drinking water standard (MCLDCA: 0.005 mg/L, ∼0.05 μM). With increasing 1,2-DCA concentration in the reactor influent, we did not observe increases in the relative abundance of nonDehalococcoides populations in 16S rRNA gene amplicon libraries. However, given that members of Desulfitobacterium, Geobacter, and Dehalobacter genera have all been shown to be capable of growth on 1,2-DCA, the molecular approaches introduced here could reveal insights into comparative fitness of Dehalococcoides and non-Dehalococcoides organohalide-respiring populations in more genetically diverse 1,2-DCA impacted environments. Given its phylogenetic specificity, the hupL-based molecular PCR assay developed in this study could also be used to track changes in Dehalococcoides population structure in samples collected in the field, even where Dehalococcoides is a rare member of the overall microbial population.

population due to differences in the affinity of these populations’ reductive dehalogenases for 1,2-DCA at micromolar concentrations. When the EV consortium was dominated by the DhcEV-TceA population, we measured a Monod half saturation parameter (KS,DCA,EV) of 127 μM (95% CI [100, 170]) (Table 3). This half saturation concentration is an order of magnitude lower than the KM,DCA,VcrA (1300 μM) reported in a recent study of heterologously expressed VcrA enzyme.20 Notably an alternative vinyl chloride reductive dehalogenase (BvcA), not abundant in the EV consortium, also has reported activity for 1,2-DCA in vitro, but the KM, DCA, BvcA has not been estimated.19 Therefore, it is difficult to predict whether VC-specialized Dehalococciodes populations expressing the bvcA rather than vcrA gene would similarly be outcompeted by TceAcontaining populations as was observed in the EV consortium. Further studies are needed to determine whether 1,2-DCA exposure will alter the population frequency of Dehalococcoides in other consortia (e.g., KB-1, SDC-9) toward the ecological dominance of TceA-type Dehalococcoides. If so, the presence of significant 1,2-DCA in cocontaminated groundwater may shift Dehalococcoides population structure away from ideal conditions for rapid and complete detoxification of vinyl chloride. In chloroethane-dominated sites, the long-term enrichment of a TceA-type population could result in preferential utilization of 1,2-DCA with the undesirable VC accumulation when electron donor becomes limiting. Our finding that 1,2-DCA degradation was strongly inhibited by cis-DCE, however, suggests that rapid 1,2-DCA dihaloelimination by TceA-containing Dehalococcoides is unlikely to occur when steady state cis-DCE concentrations remain higher than 25 μM (Figure 4). This inhibitory substrate interaction

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ASSOCIATED CONTENT

* Supporting Information S

The Supporting Information is available free of charge on the ACS Publications website at DOI: 10.1021/acs.est.6b02957. File 1: Dual-indexed barcoded oligonucleotide primer sequences used to amplify the 16S rRNA and hupL genes (PDF) File 2: Figure S1 displays the diminished VC dehalogenation activity after long-term exposure to 1,2DCA. Figure S2 shows the increase in VC dehalogenation threshold after long-term exposure to 1,2-DCA. Figure S3 presents data on the magnitude of inhibition on 1,2-DCA degradation kinetics by cis-DCE. Figure S4 shows the native-PAGE gel cutting pattern. Table S1 reports the native-PAGE gel activity and peptides detected by mass spectrometry. Details on medium preparation, reactor sampling, and blue native-polyacrylamide gel electrophesis (BN-PAGE) and liquid chromatography tandem mass spectrometry (LC-MS/ MS) are provided (PDF)

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AUTHOR INFORMATION

Corresponding Author

*E-mail: [email protected]. Notes

The authors declare no competing financial interest.

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Figure 4. Substrate interference between chloroethene and chloroethane compounds. Strong inhibitory effects (shown by solid red lines) by the chloroethane 1,1,1-TCA on Dehalococcoides mccartyi catalyzed dechlorination reactions 1 and 3 have been experimentally observed.9,14 In this study, we further observed strong inhibition (solid line) of dihaloelimination reaction 4 by cis-DCE (modeled KI, DCE→DCA, competitive model = 9 μM), and weak inhibition (dashed line) of dechlorination reaction 2 by 1,2-DCA (modeled KI,DCA→DCE, competitive model = 45 μM).

ACKNOWLEDGMENTS B.C., A.M.S., and S.H. were supported partly by NSF DMS Grant 1162538. This work was further supported by an NSF Grant (MCB-1330832). K.M.B. was additionally supported by and NSF Graduate Research Fellowship (2011103493). O.M. was supported by a Canadian Natural Science and Engineering Research Council (NSERC) PGS-D. We are grateful to Lewis Semprini and Mohammad F. Azizian for the sharing the EV-5 mixed bacterial consortium. Oxford NanoPore’s MAP early access program provided us with a MinION DNA sequencer. We also wish to thank Diego Calderon for assistance with nanopore DNA sequencing and insightful comments and Andrew Quail for help with mass spectrometry.

implies that in chloroethene-dominated environments, 1,2-DCA dihaloelimination by TceA-containing Dehalococcoides and the attendant population shift−due to a growth advantage predicated on 1,2-DCA utilization−will be slow. Moreover, the observed KS,DCA,EV (>100 μM) for 1,2-DCA indicates that the H


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1,2-Dichloroethane Exposure Alters the Population Structure

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