Article pubs.acs.org/est
Distinct Carbon Isotope Fractionation during Anaerobic Degradation of Dichlorobenzene Isomers Xiaoming Liang,† Scott O. C. Mundle,*,† Jennifer L. Nelson,‡ Elodie Passeport,† Calvin C. H. Chan,† Georges Lacrampe-Couloume,† Stephen H. Zinder,‡ and Barbara Sherwood Lollar† †
Department of Earth Sciences, University of Toronto, Toronto, Ontario M5S 3B1, Canada Department of Microbiology, Cornell University, Ithaca, New York 14853, United States
‡
S Supporting Information *
ABSTRACT: Chlorinated benzenes are ubiquitous organic contaminants found in groundwater and soils. Compound specific isotope analysis (CSIA) has been increasingly used to assess natural attenuation of chlorinated contaminants, in which anaerobic reductive dechlorination plays an essential role. In this work, carbon isotope fractionation of the three dichlorobenzene (DCB) isomers was investigated during anaerobic reductive dehalogenation in methanogenic laboratory microcosms. Large isotope fractionation of 1,3DCB and 1,4-DCB was observed while only a small isotope effect occurred for 1,2-DCB. Bulk enrichment factors (εbulk) were determined from a Rayleigh model: −0.8 ± 0.1 ‰ for 1,2DCB, −5.4 ± 0.4 ‰ for 1,3-DCB, and −6.3 ± 0.2 ‰ for 1,4-DCB. εbulk values were converted to apparent kinetic isotope effects for carbon (AKIE) in order to characterize the carbon isotope effect at the reactive positions for the DCB isomers. AKIE values are 1.005 ± 0.001, 1.034 ± 0.003, and 1.039 ± 0.001 for 1,2-DCB, 1,3-DCB, and 1,4-DCB, respectively. The large difference in AKIE values between 1,2-DCB and 1,3-DCB (or 1,4DCB) suggests distinct reaction pathways may be involved for different DCB isomers during microbial reductive dechlorination by the methanogenic cultures.
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INTRODUCTION Dichlorobenzenes (DCBs) have been extensively used in industrial and domestic products such as solvents and pesticides, and occur as chemical intermediates.1 Toxicity studies show that exposure to DCBs can adversely impact wildlife and humans, and they have therefore been listed as priority environmental pollutants by the U.S. EPA.1 Since the densities of DCB isomers are greater than water and their solubilities in water are low (from 83−147 mg/L),2 DCBs are often present as dense nonaqueous phase liquids (DNAPL) at contaminated sites. Both aerobic3,4 and anaerobic5,6 biodegradation can contribute to the natural attenuation of DCBs. Aerobic degradation of DCBs has been intensively studied in the past decade and is reviewed elsewhere.7 A number of microbial strains are capable of using DCBs as sole carbon and energy source,7 and most of these strains are members of the genera Burkholderia, Pseudomonas, or Alcaligenes.3,4,7 Aerobic degradation of DCBs is initiated via dioxygenases with the formation of dihydrodiol intermediates, which are subsequently oxidized to chlorocatechol by dehydrogenases.7 Several strains can carry out anaerobic dehalogenation of chlorinated benzenes, such as D. mccartyi strain 195,8 D. mccartyi strain CBDB1,9,10 and strain DF-1.11 Bosma et al.5 observed the anaerobic dechlorination of all three DCB isomers (1,2-DCB, 1,3-DCB, and 1,4-DCB) in the sediments collected from Rhine River, Netherlands. Ramanand et al.12 described the transformation of DCB intermediates to monochlorobenzene (MCB) during anaerobic reductive dechlorination of hexachlorobenzene (HCB), pentachlorobenzene (PCB), and 1,2,4© 2014 American Chemical Society
trichlorobenzene (1,2,4-TCB) in soil slurries under methanogenic conditions. Complete reductive dechlorination of DCBs to benzene via an MCB intermediate was recently demonstrated in methanogenic sediments collected from a historically CB-contaminated site,6 where the organism implicated in dechlorination was Dehalobacter spp.13 Compound specific isotope analysis (CSIA) has evolved as a tool to identify biodegradation and to quantify the progress of bioremediation.14 Typically, contaminants containing the light carbon isotope are degraded faster, resulting in 13C enrichment in the remaining substrate during degradation. For carbon isotope fractionation, the magnitude of εbulk depends on many factors including: the type of bond broken (e.g., C−H or C− Cl), the reaction mechanism, and the number of carbon atoms at nonreactive positions in the contaminant.15 Fractionation associated with nondegradative processes such as volatilization, sorption and dissolution tend to be negligible and generally within analytical uncertainty relative to the fractionation from bond cleavage during degradation reactions.16,17 In contrast, carbon isotope fractionation has been characterized for numerous organic contaminants, including chlorinated ethenes,18−21 and chlorinated ethanes.22,23 To date, carbon isotope fractionation of CBs has only been investigated for TCBs and MCB.24−26 Under anaerobic Received: Revised: Accepted: Published: 4844
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anaerobic chamber in order to ensure equilibrium partitioning between headspace and liquid phase. The microcosms were incubated in the dark at room temperature in the anaerobic chamber. Abiotic controls containing only deionized water and DCBs were also prepared. Each experiment was set up with duplicate bottles. A separate set of bottles was used for each of the three DCB isomers. Analytical Methods. Concentration of DCBs and of daughter products MCB and benzene were measured by manually injecting 300 μL of headspace sample from each experimental bottle into a Varian 3380 gas chromatograph (GC) equipped with a flame ionization detector (FID) and a Restek Rtx-35 capillary column (60 m × 0.53 mm × 1.5 μm). Both injector and detector temperatures were 200 °C. The initial oven temperature was 50 °C, immediately increased to 200 °C at 15 °C/min, and held isothermally at 200 °C for 1 min. Three point external calibration curves were prepared daily. Relative standard deviations for samples and standards using this method were typically ±5%. Carbon isotope analysis for CBs was conducted by direct injection of headspace samples with gastight syringes into a Hewlett-Packard 6890 GC interfaced with a combustion oven in line with a Finnigan MAT Delta plus XL isotope ratio mass spectrometer after the method of Slater et al.17 The combustion oven temperature was controlled at 980 °C. The GC was equipped with a VOCOL capillary column (60 m × 0.25 mm × 1.5 μm), and the GC temperature was set at 40 °C, immediately ramped to 210 °C at 25 °C/min, and held for 8 min. The GC injector temperature was 180 °C. Abiotic controls and DCB isomer isotope working standards were run daily. The total analytical uncertainty for δ13C was ±0.5 ‰, incorporating both reproducibility and accuracy after the method of Sherwood Lollar et al.28
conditions, large carbon isotope fractionation was observed during reductive dehalogenation of 1,2,3-TCB (εbulk = −3.4 ± 0.3‰) and 1,2,4-TCB (εbulk = −3.2 ± 0.5‰) by D. mccartyi strain CBDB1.24 Liang et al.26 confirmed these values for biodegradation of 1,2,4-TCB and also demonstrated large carbon isotope fractionation for dechlorination of MCB (εbulk = −5.0 ± 0.2 ‰) by methanogenic cultures.26 No significant isotope fractionation was found during aerobic degradation of either TCBs or MCB.24−26 Liang et al.26 suggested that the larger ε values observed for anaerobic degradation likely result from rate-limiting C−Cl bond cleavage and the smaller ε values in aerobic degradation result from rate-limiting C−O bond formation.26 Stelzer et al.27 interpreted an overall enrichment of the cumulative isotopic signature of CBs across an anoxic contaminated aquifer as supportive of some degree of biodegradation. Confirming such interpretations and potentially quantifying the degree of biodegradation however, requires further investigation of the carbon isotope effects that occur during CBs biodegradation. To date, no isotopic fractionation studies have been published on the three isomers of DCB, although there is considerable interest in the degradation of these compounds−particularly for a recent Dehalobacter-containing culture demonstrated to be capable of complete dechlorination of TCBs, DCBs, and MCB to benzene.13 The goals of this study were (1) to determine carbon isotope fractionation for the three DCB isomers during anaerobic reductive transformation by methanogenic cultures and (2) to explore the potential dechlorination mechanisms and effects of both enzyme catalysis and chlorine substituents on the isotope fractionation of DCB isomers. To our knowledge, this is the first report of εbulk values for DCB isomers, which is the key information for monitoring in situ biodegradation of DCBs by CSIA.
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RESULTS Abiotic Controls. During the course of degradation experiments, no significant variation in the concentrations of DCB isomers was observed in abiotic controls, and the δ13C values in control bottles remained within ±0.5 ‰ of known δ13C values for the DCB isomer isotopic working standards for all experiments (data not shown). Anaerobic Dehalogenation of DCBs in Methanogenic Microcosms. Plots of concentration versus time for biodegradation of all three DCB isomers are shown in Figure 1. At time zero, background benzene (20−50 μM) was found in all microcosms (Figure 1), reflecting benzene remaining in the sediments used to set up the microcosms. During the biodegradation experiments sequential dechlorination occurred for all three DCB isomers (Figure 1A−C). Each of the three DCB isomers was transformed to MCB followed by benzene. Benzene from the dechlorination of MCB started to build up significantly only when each individual DCB isomer was close to complete consumption as evidenced by the fast accumulation of benzene at the end of experiments (Figure 1B,C). Complete degradation occurred in ∼50 h for 1,2-DCB (Figure 1A), ∼110 h for 1,3-DCB (Figure 1B), and ∼275 h for 1,4DCB (Figure 1C). Observed first-order rate constants were calculated from nonlinear regression fitting to the integrated first-order rate expression for 1,2-DCB (kobs = 0.84 ± 0.08 d−1), 1,3-DCB (kobs = 0.40 ± 0.04 d−1), and 1,4-DCB (kobs = 0.12 ± 0.01 d−1). Carbon Isotope Fractionation during Biodegradation of DCBs. Plots of δ13C and fraction remaining during
EXPERIMENTAL SECTION Chemicals. The following chemicals were obtained from Sigma-Aldrich (St. Louis, MO): 1,2-DCB (99%), 1,3-DCB (99%), 1,4-DCB (99%), MCB (99.5%), benzene (99.8%), acetone (99%), bicarbonate (99.5%), yeast extract. Microcosm Setup. Microbial cultures derived from methanogenic sediments were collected from an anoxic zone at the DuPont Chambers Works site (Salem County, NJ), which has been historically contaminated by CBs and anilines.6,26 Microcosm bottles for this study were set up in duplicate in 120 mL glass bottles capped with Mininert valves. Microcosms containing 20 g of wet sediment and 50 mL of oxygen-free deionized water were prepared in an anaerobic glovebox (Labconco Co., Kansas City, MO). Each microcosm was inoculated with 2 mL slurry from prior actively reductively dechlorinating microcosms. Sodium bicarbonate (1 g/L) and yeast extract (0.2 g/L) were added to the microcosms as a buffer and an electron donor, respectively. The headspaces of the experimental bottles consisted of 80% N2 and 20% CO2. The bottles were spiked with appropriate amounts of concentrated stock solution (1 M) that had been prepared in acetone (1,2-DCB: 30 μL; 1,3-DCB: 30 μL; 1,4-DCB: 44 μL) to produce starting concentrations between 200 and 350 μM. The DCB isomers had been previously isotopically characterized and had δ13C values of −28.5‰ ± 0.5‰, −23.6‰ ± 0.5‰, −24.8‰ ± 0.5‰ for 1,2-DCB, 1,3-DCB, and 1,4-DCB, respectively (Supporting Information (SI) Figure S1A−C). Experimental bottles were manually shaken for 30 min in the 4845
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Figure 2. Carbon isotope fractionation during anaerobic reductive dechlorination of (A) 1,2-DCB, (B) 1,3-DCB, (C) 1,4-DCB. Error bars on f represent ±7% total uncertainty. Error bars on δ13C represent ±0.5‰ (reproducibility and accuracy). The solid line represents a Rayleigh model fit. In all cases data from both duplicate bottles from each experiment are plotted.
Figure 1. Anaerobic reductive dehalogenation of individual DCB isomers by methanogenic cultures in laboratory microcosms (A) 1,2DCB, (B) 1,3-DCB, (C) 1,4-DCB. Error bars on C represent ±5%. Results are shown for one of the duplicate bottles for each DCB isomer. In all cases both duplicates showed similar results.
values are −0.8 ± 0.1‰ for 1,2-DCB, −5.4 ± 0.4‰ for 1,3DCB, and −6.3 ± 0.2‰ for 1,4-DCB (Table 1).
transformation of each DCB isomer in both duplicate bottles are shown in Figure 2A−C. Compared to the δ13C0 of −28.1 ‰, the carbon isotope signature of 1,2-DCB was enriched in 13 C by just over 2‰ to −26.0‰ when more than 95% of the substrate was degraded (Figure 2A). In comparison, δ13C values of 1,3-DCB and 1,4-DCB became significantly enriched in 13C during the course of anaerobic degradation (Figure 2B,C). At the end of the experiments, the overall changes in δ13C for 1,3DCB and 1,4-DCB were approximately 14.5‰ and 16.0‰, respectively. In all cases, carbon isotope fractionation of each of the DCB isomers fit the Rayleigh model (Figure 2A−C, and SI Figure S2A−C) with R2 values >0.96 (Table 1). The εbulk values in duplicate bottles for each DCB isomer were statistically the same within 95% confidence intervals (Table 1), and hence an overall εbulk value for each isomer was calculated for both duplicates combined.29 Table 1 contains both the ε values for individual duplicates and for both combined. The overall εbulk
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DISCUSSION Different Dehalobacter strains have been implicated in anaerobic dechlorination of 1,2-DCB, 1,3-DCB, and 1,4-DCB to chlorobenzene (CB).13 Nelson et al.13 isolated 1,2-DCB degrading Dehalobacter sp. strains that did not degrade 1,3DCB or 1,4-DCB. They also isolated 1,3-DCB and 1,4-DCB degrading Dehalobacter sp. strains that did not effectively degrade 1,2-DCB, and an enrichment culture capable of dechlorinating para-substituted chlorobenzenes.30 Measured εbulk values from sediment cultures containing all Dehalobacter strains for dechlorination of 1,2-DCB, 1,3-DCB, and 1,4-DCB were −0.8‰, −5.4‰, and −6.3‰, respectively. The εbulk values derived from the Rayleigh model represent the extent of carbon isotope fractionation on the basis of the whole molecule for each DCB isomer. In order to characterize the isotope effect 4846
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Table 1. Measured carbon bulk enrichment factors (εbulk) and apparent kinetic isotope effects for carbon (AKIEC) for reductive dehalogenation of DCB isomers compound
replicate no.
1,2-DCB
1 2 overall 1 2 overall 1 2 overall
1,3-DCB
1,4-DCB
ε (‰)a −0.8 −0.7 −0.8 −5.4 −5.5 −5.4 −6.0 −6.5 −6.3
± ± ± ± ± ± ± ± ±
R2
0.1 0.1 0.1 0.4 0.4 0.4 0.3 0.4 0.2
0.97 0.97 0.96 0.97 0.97 0.97 0.99 0.99 0.99
nb
xb
zb
6
2
2
6
2
2
6
2
2
AKIECc 1.005 1.004 1.005 1.034 1.034 1.034 1.037 1.041 1.039
± ± ± ± ± ± ± ± ±
0.001 0.001 0.001 0.003 0.003 0.003 0.002 0.003 0.001
a
Enrichment factors are determined from the slope of the linear regression of ln f versus ln (R/R0), and uncertainties represent 95% confidence intervals calculated from standard deviation of regression slope. bSee text Discussion. cUncertainties are 95% confidence intervals calculated from propagation of error.
Figure 3. Summary of reaction mechanisms for cobalt corrinoid-dependent reductive dechlorination of chlorinated ethenes by inner-sphere electron transfer via (A) nucleophilic addition, (B) nucleophilic substitution, and by outer-sphere electron transfer via (C) single electron transfer.
Although pathway-dependent kinetic isotope effects are commonly invoked to explain large differences in observed AKIE’s, there are a limited number of reasonable catalytic mechanisms that can be used by the enzyme to convert these DCB isomers to the same chlorobenzene product. The different reactivity patterns and isotope fractionation observed in these degradation reactions can be used to provide additional insight into the possible pathway(s) involved in these dechlorination reactions. Mass Transfer Effects. Rate limitations on mass transfer for contaminants across both phase boundaries31 and cell membranes32 (in whole-cells microcosm experiments) have been shown to suppress the magnitude of AKIE’s. The concentrations of the DCB-isomers were below saturation limiting this system to two-phases (headspace and solution phase).2 The importance of mass transfer limitations can be predicted based on dimensionless Henry’s Law constants; however, a significant effect was not expected between the DCB-isomers since dimensionless Henry’s law constants are very similar for these isomers (1,2-DCB = 0.08, 1,3-DCB = 0.12, and 1,4-DCB = 0.10).33,34 Membrane transport effects
at the reactive position of the DCB isomer, apparent kinetic isotope effect values for carbon (AKIEC) were calculated according to the following equation as described in Elsner et al.:15 z ·n·εbulk 1 = +1 AKIEC x· 1000
(1)
where n is the number of C atoms in the molecule, x is the number of C atoms with the potential for bond cleavage, and z is the number of C atoms having equal reactivity. All values of n, x, and z and the resulting values of AKIEC are summarized in Table 1. The AKIEC values were similar for 1,3-DCB (1.034 ± 0.003) and 1,4-DCB (1.039 + 0.001), but nearly an order of magnitude larger for 1,2-DCB (1.005 ± 0.001) (Table 1). The large difference in AKIEC for the dechlorination of 1,2DCB relative to 1,3-DCB/1,4-DCB likely corresponded to the different Dehalobacter strains responsible for degradation.13 Ultimately, it is not known whether similar enzymes with different substrate specificity or different degradation pathways are responsible for dechlorination of each DCB isomer. 4847
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have been reported for PCE dechlorination;32 however, this phenomenon was not observed to alter the AKIE’s for both 1,1,1-TCA and 1,1-DCA.23 Ultimately, these effects have been shown to be relatively small to date and even if they were present to some degree in this study they cannot account for the order of magnitude difference in AKIE’s observed between 1,2-DCB and 1,3-DCB/1,4-DCB.35 These effects would be expected to have a greater impact on the AKIE’s for different contaminants with different chemical properties, rather than for constitutional-isomers of the same contaminant with similar chemical properties. The large difference in AKIE for 1,2-DCB and 1,3-DCB/1,4-DCB is more consistent with differences in commitment factors or catalysis at the level of the enzyme in these organisms. Insight into these factors can be gained by considering DCB reductive dechlorination within the context of the reported mechanisms for reductive dechlorination of chlorinated ethenes. Reductive Dechlorination: Chlorinated Ethenes. Anaerobic microbial degradation of chlorinated ethenes generally involves cobalt corrinoid-dependent enzymes. Several different routes have been proposed to account for catalysis including inner-sphere electron transfer processes (via nucleophilic addition, or nucleophilic substitution), and outer-sphere single electron transfer mechanisms (Figure 3).36 The general mechanism for nucleophilic addition involves formation of a carbon−cobalt bond and addition of hydrogen, generating a cofactor-alkane enzyme-bound intermediate that subsequently leads to elimination of the chlorine and cofactor to form the singly dechlorinated product (Figure 3A). Nucleophilic substitution involves direct substitution of a carbon-chlorine bond with a carbon−cobalt bond leading to a cofactor-alkene enzyme-bound intermediate followed by substitution of the cofactor for hydrogen to form the product (Figure 3B). Single electron transfer routes are initiated by a reaction between the cofactor and chlorinated substrate that lead to the formation of a radical ion intermediate that can either be protonated to form the product, or react with the cofactor to form a carbon−cobalt bond with subsequent product-forming steps analogous to latter part of the nucleophilic substitution mechanism (Figure 3C).37 Currently, there is no direct evidence implicating a cobalt corrinoid-cofactor in the DCB reductase; however, it is likely that cofactors are also involved in the Dehalobacter DCB degrading strains used in this study as they have been implicated in PCE dechlorination by Dehalobacter restrictus.38 Therefore, if the fundamental difference in reactivity between chlorinated alkenes and chlorinated benzenes is accounted for, the enzyme-catalyzed routes implicated in the reductive dechlorination of chlorinated ethenes can provide a basis for interpreting the potential cofactor-dependent mechanisms most likely involved in the reductive dechlorination of chlorinated benzenes. Reductive Dechlorination: Chlorinated Benzenes. Cofactor-catalyzed reductive dechlorination of substituted benzenes can be rationalized based on three fundamental mechanistic scaffolds: (1) electrophilic aromatic substitution39 (EAS; Figure 4A), (2) nucleophilic aromatic substitution40 (NAS; Figure 4B), and (3) homolytic aromatic substitution41 (HAS; Figure 4C). In the dechlorination of DCB-isomers, EAS and NAS mechanisms proceed via initial rate-limited addition of an electrophile (proton) or a nucleophile (metal-cofactor) leading to a cationic or anionic resonance-stabilized intermediate that subsequently breaks a carbon−chlorine bond to form chlorobenzene (Figure 4A,B). The analogous nucleophilic
Figure 4. Summary of possible reaction mechanisms for dechlorination of DCB isomers to chlorobenzene by inner-sphere electron transfer via (A) EASelectrophilic aromatic substitution, (B) NAS nucleophilic aromatic substitution, and outer-sphere electron transfer via (C) HAShomolytic aromatic substitution (single electron transfer). Although these mechanism have been adapted from cobalt corrinoid-dependent reductive dechlorination of chlorinated ethenes the identity of the cofactor has been left anonymous (designated “M”) because cobalt corrinoid-dependent enzymes have not been directly implicated in DCB dechlorination.
addition mechanism for chlorinated ethenes, where the cofactor and a hydrogen are added across the double bond can be adapted to aromatic reactions as either an EAS or NAS mechanism depending on the order of bond formation and cleavage. The EAS mechanism (Figure 4A) would involve initial addition of a proton (electrophile) to the aromatic ring followed by addition of the cofactor (nucleophile). Depending on the binding orientation of the substrate, if the cofactor is directed at the carbon adjacent to the reaction-site, it can provide a driving force for asynchronous protonation and addition of the cofactor. The NAS mechanism (Figure 4B) is comparable to both inner-sphere mechanisms for chlorinated ethenes (nucleophilic addition and nucleophilic substitution), which would involve addition of the cofactor (nucleophile) to the aromatic ring followed by elimination of chlorine. Figure 4B represents the analogous nucleophilic substitution mechanism for chlorinated ethenes; nucleophilic addition would involve protonation of the aryl-carbanion followed by loss of the proton and elimination of chlorine that ultimately generates the same cofactor-conjugated enzyme-bound intermediate. HAS is directly analogous to the outer-sphere single electron transfer mechanism for chlorinated ethenes and will initially form a benzyl-radical ion intermediate upon dissociative release of chlorine (Figure 4C). Although dissociative reactions on aromatic rings are less common, this mechanism has been extensively studied in the abiotic decomposition of aromatic diazonium salts, where a labile carbon−nitrogen bond is cleaved in the first step releasing N2.41 Upon rate-limited release of N2 the benzyl-cation/radical ion rapidly reacts to form the aromatic substitution product. Figure 4C shows a representation of these routes within the context of 4848
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significant isotope fractionation was attributed to rate-limiting C−O bond formation. 26 Since anaerobic and aerobic degradation of chlorinated benzenes lead to different reaction products, pathway-dependent kinetic isotope effects are reasonable. In this case, comparison of different DCB-isomers that lead to the same product, where these differences likely arise from similar enzyme-catalyzed reactions, can provide additional insight into the factors that influence the expression of AKIE’s and pathway-dependent kinetic isotope effects. Brown and Drury48 reported the nitrogen kinetic isotope effects (AKIEN) for the abiotic decomposition of 2-methylbenzenediazonium (1.045 ± 0.001), 3-methylbenzenediazonium (1.047 ± 0.001), and 4-methylbenzenediazonium (1.047 ± 0.001). These reactions are analogous to dechlorination of the DCB-isomers via a HAS mechanism. The AKIEN was not affected by the position of the chlorine substituent on this ring, which is consistent with the hypothesis that the magnitude of AKIE’s will remain relatively constant for similar transitionstates despite changes in reaction rates. The large magnitude of the AKIEN suggests C−N bond-breaking is rate-limited in this reaction, which is comparable to the large AKIEC observed for carbon in the dechlorination of 1,3-DCB and 1,4-DCB. These AKIE’s can be generated in HAS mechanisms, where dissociative loss of chlorine leads to an aryl-radical ion generated upon single electron transfer and carbon−chlorine bond cleavage (Figure 4C). Alternatively, the AKIEC for 1,2DCB (1.005) supports a NAS mechanism, where carbonchlorine bond cleavage is not a significant part of the ratelimiting transition-state. Since dechlorination of 1,2-DCB likely proceeds via a NAS mechanism, a logical hypothesis that can account for the different AKIE’s is that the enzyme responsible for dechlorination of 1,3-DCB/1,4-DCB may proceed via the HAS single electron transfer mechanism. An alternative interpretation for the different AKIEs between 1,2-DCB and 1,3-DCB/1,4-DCB, based on the theoretical framework for commitment to catalysis developed by Northrop49 implicates the possibility that an NAS mechanism could be involved for all three DCB-isomers. In this case, the change in AKIE between 1,2-DCB and 1,3-DCB/1,4-DCB would arise from a combination of factors that affect the reversibility50 of the addition step for carbon−cofactor bond formation (if addition is reversible C−Cl bond cleavage becomes part of the rate expression), and the influence of secondary isotope effects51 at nonreactive positions. However, it is not well understood if the impact of these effects can lead to the observed order of magnitude difference in AKIE’s for a single constitutional isomer among the series of DCB-isomers examined in this study. Therefore, it is more likely that the different AKIE’s for reductive dechlorination of the DCBisomers reflects the underlying mechanisms (NAS and/or HAS), which are influenced by similar factors implicated in regulating the mechanisms for reductive dechlorination of chlorinated ethenes (outer sphere versus inner sphere electron transfer).36,37 Environmental Significance. As many as 20% of sites on the National Priority List in the United States have been identified as contaminated by at least one of the DCB isomers. Their widespread detection has led to increased public concern and the need for development of efficient remediation strategies and tools. As noted in the U.S. EPA Guide for CSIA,14 the establishment of enrichment factors in the laboratory plays a key role in the ability to quantify the extent of in situ biodegradation (B, %) using carbon isotope ratios of
DCB dechlorination, where initial dissociation of the carbon−chlorine bond via inner-sphere single electron transfer is the rate-limiting step. Catalytic Mechanisms and Structure−Reactivity Models. Aromatic substitution reactions follow a classic pattern, where the reaction rates are influenced by substituent effects.42 For the DCB-isomers, differences in reaction rates will arise from the inductive electron withdrawing effect of the chlorine substituents on the benzene ring. The proximity of substituents in the 1,2-isomer will produce the most significant effect. The observed electronic substituent effect decreases as the distance between the substituent and the reaction-site increases in 1,3DCB and 1,4-DCB.43 The electron withdrawing effect of the chlorine atom will increase the observed rate for NAS and HAS mechanisms through a combination of transition-state stabilization and ground-state destabilization for rate-limited addition of the metal cofactor or benzyl-radical, which lowers the kinetic energy barrier for nucleophilic addition and dissociation of chlorine via single electron transfer.41 In this case, the fastest reaction rate is expected from the 1,2-DCB isomer. Alternatively, in EAS reactions where positive charge is developed in the transition-state, electron withdrawal will increase the kinetic energy barrier for addition of the electrophile to the aromatic ring and 1,2-DCB would be expected to degrade with the slowest observed rate. Abiotic dechlorination provides a model system where the influence of electronic substituent effects on the reactivity patterns and reaction rates can be predicted by the structure− reactivity relationships between each of the DCB isomers. Based on reported reaction rates for aromatic substitution of a chlorine atom for hydrogen with a nickel/activated carbon catalyst, the reactivity order should follow the classic pattern expected for a NAS mechanism (1,2-DCB > 1,3-DCB > 1,4DCB).44 In this study, consistent with previously reported results by Fung et al.,6 anaerobic dechlorination by Dehalobacter for 1,2-DCB was complete in ∼50 h (kobs = 0.84 ± 0.08 d−1), followed by 1,3-DCB (∼100 h, kobs = 0.40 ± 0.04 d−1) and 1,4DCB (∼250 h, kobs = 0.12 ± 0.01 d−1), which is consistent with a NAS or HAS mechanism, where the differences in observed reaction rates can be partially attributed to substituent effects. Since the EAS mechanism is expected to produce the opposite reactivity pattern (1,4-DCB > 1,3-DCB > 1,2-DCB) by destabilizing the transition-state for addition of an electrophile to the aromatic ring, it is the least likely catalytic route employed by Dehalobacter catalyzed dechlorination in this system. Although the reactivity pattern and relative differences in reaction rates limit the possible reaction mechanisms to NAS and/or HAS, a closer look at the origin of kinetic isotope effects and the factors that influence the expression of AKIE’s is needed to account for the large difference in the magnitudes of the AKIEC between 1,2-DCB and 1,3-DCB/1,4-DCB. Kinetic Isotope Effects. Large changes in AKIEC for processes taking place via similar mechanisms are not typically expected, as it has been shown that significant changes in transition-state bonding only produce small effects on the magnitude of an AKIEC.45,46 Therefore, large differences in isotope fractionation are commonly attributed to pathwaydependent kinetic isotope effects, where different biocatalyzed processes are implicated.23,47 For example, the large carbon isotope fractionation observed in anaerobic dechlorination of 1,2,4-TCB was attributed to rate-limited carbon−chlorine bond cleavage, in contrast to aerobic degradation, where less 4849
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field samples based on the following equation: B = (1 − f) × 100 where an expression for f is derived from the Rayleigh equation. The present study demonstrated large carbon isotope fractionation during reductive dehalogenation of 1,3-DCB and 1,4-DCB in methanogenic microcosms, where Dehalobacter spp. is the major class of organisms acting on dechlorination. These findings suggest that field-derived carbon isotope signatures of 1,3-DCB and 1,4-DCB can serve as both qualitative and quantitative indicators of anaerobic biodegradation of these compounds in contaminated groundwater. The small isotopic fractionation (
