Carbon Isotopic Fractionation during Aerobic Vinyl Chloride

Vinyl chloride (VC), a common contaminant in North American groundwaters (1), is a ... The biochemical pathway of aerobic VC-assimilation is currently...
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Environ. Sci. Technol. 2005, 39, 1064-1070

Carbon Isotopic Fractionation during Aerobic Vinyl Chloride Degradation MICHELLE M. G. CHARTRAND,† ALISON WALLER,‡ T I M O T H Y E . M A T T E S , §,| MARTIN ELSNER,† GEORGES LACRAMPE-COULOUME,† JAMES M. GOSSETT,§ ELIZABETH A. EDWARDS,‡ AND B A R B A R A S H E R W O O D L O L L A R * ,† Stable Isotope Laboratory, Department of Geology, University of Toronto, Toronto, Ontario, Canada M5S 3B1, Department of Chemical Engineering and Applied Chemistry, University of Toronto, Toronto, Ontario, Canada M5S 3E5, and School of Civil and Environmental Engineering, Cornell University, Ithaca, New York 14853

Vinyl chloride (VC) is a carcinogenic contaminant commonly found in groundwater. Much research has focused on anaerobic reductive dechlorination of VC, and recently on aerobic VC degradation. In this study, the stable carbon isotope enrichment factor associated with aerobic VC assimilation was determined for Mycobacterium sp. strains JS60, JS61, and JS617 and Nocardioides sp. strain JS614. The enrichment factors ranged from -8.2 ( 0.1 to -7.0 ( 0.3 ‰ and did not change as a function of biomass concentration. The measured enrichment factors for aerobic VC degradation were smaller than those reported for anaerobic VC degradation. Enrichment factors can also be expressed in terms of kinetic isotope effects (KIEs), 12k/13k, which result from the difference in reaction rates of bonds containing light and heavy isotopes. The KIEs for aerobic VC degradation (1.01 ( 0.001) were smaller than those for anaerobic VC degradation (1.03 ( 0.007). From the perspective of bond breakage during a chemical reaction, the larger KIE associated with anaerobic VC degradation as compared to aerobic VC degradation agrees with KIE theory. This theory predicts that larger fractionations can be expected in reactions where heavier atoms are involved (i.e., C-Cl bond for anaerobic versus CdC for aerobic) and in reactions involving large changes in vibrational frequencies of the molecule between its ground state and transition state (i.e., C-Cl cleavage versus CdC epoxidation). The significant fractionation observed during aerobic VC degradation suggests that stable carbon isotope measurements may be used as a tool to distinguish between biodegraded and nonbiodegraded VC.

* Corresponding author phone: (416)978-0770; fax: (416)978-3938; e-mail: [email protected]. † Department of Geology, University of Toronto. ‡ Department of Chemical Engineering and Applied Chemistry, University of Toronto. § Cornell University. | Present address: Dept of Civil and Environmental Engineering, University of Iowa, Iowa City, IA 52242. 1064

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Introduction Vinyl chloride (VC), a common contaminant in North American groundwaters (1), is a concern due its toxicity and carcinogenic properties (2). VC contamination in groundwater is primarily a result of incomplete reductive dechlorination of more highly chlorinated ethenes (3, 4) and ethanes (5, 6). VC may be degraded under both anaerobic and aerobic conditions. Anaerobically, VC can be oxidized to CO2 and CH4 (7) or reductively dechlorinated to ethene (ETH) (3, 4). Complete, sequential reductive dechlorination of tetrachloroethene (PCE) and trichloroethene (TCE) to ETH has been documented at contaminated sites (8, 9). To date, only two microorganisms capable of reducing PCE and TCE to ETH have been isolated. Dehalococcoides etheneogenes (DHE) Strain 195 reduces PCE to ETH, although the final step, reduction of VC to ETH, is slow, and not exploited for growth (4). Strain BAV-1, on the other hand, is able to utilize VC and all dichloroethene (DCE) isomers as electron acceptors for growth, with cometabolic reductions of PCE and TCE to cis1,2-dichloroethene (cis-DCE) (3). Microorganisms capable of complete reduction of PCE and TCE to ETH are not always present at contaminated sites, and even when they are present, many site-specific factors may contribute to the accumulation of VC as the end product of dechlorination (10, 11). At some sites, anaerobic VC plumes may migrate into downgradient aerobic groundwater zones, where VC is subject to aerobic oxidation processes. Other carbon sources, such as ethane, ethene, benzene, toluene, ethylbenzene, and xylene, (BTEX) may be present or migrate with the VC plume. Aerobic bacteria growing on these compounds express relatively nonspecific oxygenases that fortuitously degrade VC (12, 13). In the absence of cosubstrates, other aerobic bacteria may use VC as a sole carbon and energy source. For many years, the only known VC-assimilating bacteria included four Mycobacterium strains (14, 15) and two Pseudomonas strains (16, 17), although it was unclear if these isolates were representative of the phylogenetic and kinetic diversity of VC assimilating bacteria in the environment. Recently, 12 VC-assimilating strains were isolated from environmental samples (11 Mycobacterium strains and one Nocardioides strain), indicating that VC assimilating bacteria are widespread and may play a major role in aerobic natural attenuation of VC (18). Despite evidence that VC is readily oxidized under aerobic conditions, and the widespread distribution of VC assimilating bacteria in the environment, aerobic VC oxidation in the field remains difficult to prove. Disappearance of VC may be due to dilution or sorption, and there are no specific geochemical markers to distinguish VC degradation from other degradation processes. The biochemical pathway of aerobic VC-assimilation is currently poorly understood (Figure 1). The initial step in VC assimilation is catalyzed by an alkene monooxygenase (AkMO), which forms the corresponding epoxide chlorooxirane. All known VC-assimilating bacteria also appear to use AkMO to form epoxyethane during growth on ETH. Epoxyalkane:Coenzyme M transferase (EaCoMT) catalyzes the first step in chlorooxirane and epoxyethane metabolism in Mycobacterium sp. strain JS60 (19). EaCoMT transfers a molecule of CoM (2-mercaptoethanesulfonate) to epoxyethane, opening the epoxide ring, and forming a 2-hydroxyethyl-CoM conjugate (20). EaCoMT also catalyzed the degradation of chlorooxirane in the presence of CoM, but the expected intermediate was not observed. Until recently, CoM was considered an exclusively methanogenic cofactor, 10.1021/es0492945 CCC: $30.25

 2005 American Chemical Society Published on Web 01/06/2005

FIGURE 1. Proposed pathway of VC assimilation in Mycobacterium sp. strain JS60. Compounds in brackets have not been identified, and dotted arrows represent hypothetical reactions (modified from Coleman et al. (19)). although it is now evident that CoM is commonly used by bacteria carrying out aliphatic alkene oxidation (20). Compound specific isotope analysis (CSIA) has been shown to provide a powerful means of identifying chlorinated ethene and ethane biodegradation in both laboratory and field studies (5, 21-29). For chlorinated ethenes, processes such as dissolution, volatilization, and sorption do not significantly affect the ratios of stable carbon isotopes at equilibrium (30-34). In situations where extensive removal of mass due to vaporization has taken place (i.e., soil vapor extraction or bioventing), the effect of these small fractionation factors will be maximized and some degree of isotopic shift might be measurable (31, 34). However, the total fractionation effect is small as compared to the large fractionation measured during biodegradation for chlorinated ethenes (31), especially if mass loss due to volatilization is not a major process at the site in question. Isotopic fractionation occurs due to differing rates of reaction of each isotopic species (35, 36). The heavier isotopic species tend to form stronger bonds than the lighter isotopic species, resulting in faster reaction rates for molecules containing the lighter isotope as compared to molecules containing heavier isotopes, particularly in biological systems. For instance, in aerobic VC degradation (reported in this paper), as VC is oxidized to form chlorooxirane, the isotope value of the remaining VC should become more enriched in 13C (i.e., a less negative value) due to the preferential incorporation of 12C molecules into the product. To date, little is known about fractionation factors for aerobic VC degradation. The main objective of this study was to determine the fractionation factors for three Mycobacterium strains and a Nocardioides strain isolated by Coleman et al. (18). Kinetic studies demonstrated a significantly larger growth yield for the Nocardioides strain as compared to the other three Mycobacterium strains, suggesting a potentially different biochemical pathway. It was hypothesized that, if the VC assimilation pathways were different, this may be reflected in different fractionation factors. The fractionation factors associated with anaerobic biodegradation of VC are well established (5, 25, 26) and have been used to evaluate VC biodegradation in the field (21, 23, 28). Determining the fractionation factors associated with aerobic VC degradation will facilitate similar applications in aerobic plumes.

Experimental Section Media and Culture Bottle Preparation. All chemicals were reagent grade unless otherwise noted. The laboratory isotopic working standard VC (99.5%, Fluka) used in all experiments had previously been characterized and has a known isotope

value of -28.4 ( 0.5 ‰. Stock cultures of Mycobacterium sp. strain JS60 (ATCC BAA-494), Mycobacterium sp. strain JS61 (ATCC BAA-495), Mycobacterium sp. strain JS617 (ATCC BAA497), and Nocardioides sp. strain JS614 (ATCC BAA-499) were grown in 100 mL of minimal salts medium (MSM), initially containing approximately 150 µmol of filter-sterilized isotopic working standard VC (see above) per 260 mL bottle. MSM (18) was prepared in 1 L of water and contained the following chemicals: 0.95 g of K2HPO4, 2.27 g of KH2PO4, 0.67 g of (NH4)2SO4, and 2 mL of trace metal solution (6.37 g of EDTA, 60 g of MgSO4‚7H2O, 1 g of ZnSO4‚7H2O, 0.5 g of CaCl2‚2H2O, 2.5 g of FeSO4‚7H2O, 0.1 g of NaMoO4‚2H2O and CuSO4‚ 5H2O, 0.2 g of CoCl2‚6H2O, and 0.52 g of MnSO4‚H2O in 1 L of water). Filter-sterilized oxygen (10 mL, 99.3% (Scott)) was added to each stock culture after approximately 300 µmol of VC was degraded. Stock cultures were fed at least twice before transfer to culture bottles for isotopic experiments. All stock and culture bottles were incubated aerobically at room temperature on a rotating shaker at 50 rpm. Experimental Setup. Cultures used for isotopic analysis were prepared by adding approximately 160-186 µmol of laboratory isotopic working standard VC to 100 mL of MSM in sterile 260 mL bottles (I-Chem) capped with a mininert (Supelco) screw cap. VC was allowed to equilibrate for 1 day prior to analysis. JS60 cultures were prepared in triplicate, and JS61, JS617, and JS614 cultures were prepared in duplicate. Time-zero measurements for concentration and carbon isotope values were taken immediately after a 15 mL aliquot of culture from a stock culture was added to the experimental culture. VC concentrations and isotopic values were measured throughout the course of biodegradation. To study the effect of different initial biomass concentrations on carbon isotopic fractionation, a third JS617 replicate was prepared with a 4-fold increase in initial biomass concentration. This bottle was prepared by centrifuging 60 mL of culture from a stock culture, resuspending the pellet in 15 mL of MSM, and adding the 15 mL aliquot to the experimental culture bottle as described above. Media controls contained MSM and 143-163 µmol of VC/bottle. Killed controls contained MSM, 15 mL of JS61 or JS614 stock culture, 1.5 mL of 5% HgCl2, and 0.75 mL of 5% NaN3, and 162-184 µmol of VC/bottle. To ensure that the concentration and isotopic composition of VC were not affected by the experimental set up, sampling, or analytical methods, each control was measured throughout the experiment. Analytical Methods. VC concentrations were measured by removing a 300 µL headspace sample using gastight syringes after Slater et al. (33) and injecting the sample onto a Varian 3400 gas chromatograph (GC) with a flameionization detector, employing a 30 m × 0.53 mm i.d. GSQ (J&W Scientific) column. The temperature program was 35 °C (1 min), increased at 15 °C min-1 to 100 °C, where it was held for 3 min. The injector and detector temperatures were 200 and 220 °C, respectively, and the helium carrier gas flow rate was 7 mL min-1. VC concentrations were determined using external standards with a linear calibration (R2 > 0.99). Initial and final VC concentrations for all replicate bottles are reported in Table 1. The initial biomass concentrations were estimated using protein and optical-density (OD, 595 nm) measurements. Initial standard curves were constructed by correlating protein (assumed to comprise 55% of the cell’s dry weight (37)) to OD, and then subsequent concentrations were determined from OD. Protein concentrations were determined according to the Bradford method (38) using a colorimetric microassay kit (Bio-Rad, Hercules, California). Briefly, biomass was collected from 1 to 5 mL of culture by centrifuging at 14 000g for 20 min. The supernatant was discarded and the compacted cells were resuspended in 700 VOL. 39, NO. 4, 2005 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

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TABLE 1. Summary of Initial Protein Concentration, Initial and Final VC Concentrations for Each Replicate Bottle, Coefficient of Determination for the Rayleigh Linear Regression (R2) Used To Calculate the Enrichment Factors (E), and E for the Four Aerobic VC-Degrading Strainsa microbial strain

initial protein concn (µg mL-1)

JS60

1.3 ( 0.1

JS61

0.65 ( 0.05

JS617

0.65 ( 0.05

JS614

2.6 ( 0.1 0.65 ( 0.05

initial VC concn for each replicate bottle (µmol/ bottle)

final VC concn for each replicate bottle (µmol/ bottle)

coeff of determination for Rayleigh linear regression (R2)

E (‰)b

166 160 165 166 181 186 181 175 163 161

16 11 15 9 11 12 69 15 18 18

0.999

-8.2 ( 0.1

0.995

-7.1 ( 0.2

0.996

-7.0 ( 0.3

0.986 0.998

-7.1 ( 0.4 -7.6 ( 0.1

a Errors on  are calculated from linear regression at a 95% confidence interval. b Range of published  values for anaerobic reductive dechlorination for VC is -21.5 to -31.1‰ (5, 25, 26).

µL of 0.66 N NaOH. The cells were then incubated for 3 h. Solids were collected by centrifuging for 5 min at 5000g. The supernatant containing the protein was transferred to another vial, and 200 µL of 2 N HCl was added. Subsequently, 200 µL of dye reagent (Bio-Rad, Hercules, California) was added, and the OD was read at 595 nm. A standard curve was prepared each time using Bovine Serum Albumin (BSA) as the standard for JS60, JS61, JS617, and JS614 (R2 ) 0.979, 0.988, 0.984, and 0.988, respectively). Initial protein concentrations are reported in Table 1. CSIA measures the ratio of heavy and light elements (R ) 13C/12C) for an individual compound as compared to an international standard (V-PDB for carbon) where

δ13Ccompound ) ((Rcompound/Rstandard) - 1) × 1000 (1) The δ13C value is expressed in per mil (‰) units. Carbon isotopic values for VC were measured by removing headspace samples (100-1000 µL) after Slater et al. (33) and injecting the samples onto a Varian 3400 GC (Poraplot U column, 30 m × 0.25 mm i.d.; Chrompack; flow 1.6 mL min-1) interfaced with a combustion oven and a Finnigan MAT 252 mass spectrometer. The combustion oven consists of a copper oxide, platinum, and nickel oxide wire held at 980 °C. The chlorinated ethene is oxidized to CO2 and water. The water is removed via a Nafion membrane water trap, and CO2 enters the mass spectrometer for isotopic analysis. The GC temperature program started at 40 °C and increased at a rate of 10 °C min-1 to 185 °C. The analytical error of carbon isotopic measurements is (0.5‰ (39, 40), which incorporates both the accuracy of the measurement with respect to international standards and the reproducibility on replicate measurements of the sample. VC isotopic standards run throughout each experiment agreed with the laboratory working standard for VC (-28.4 ( 0.5‰). For each sampling event, samples for concentration and isotopic analysis were removed from the culture bottle at the same time. For many chlorinated hydrocarbons, the extent of fractionation during the initial transformation step of a contaminant can be described using the Rayleigh model equation (5, 22, 25-27, 29):

R/R0 ) f (R-1)

(2)

where R is the isotopic measurement (13C/12C) of the substrate at any given fraction remaining ( f ), R0 is the initial isotopic ratio, and R is the fractionation factor. The fractionation factor is a measure of the extent of fractionation occurring during a reaction, and if the reaction fits a Rayleigh model, R remains constant throughout the reaction (41). The fractionation 1066

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factor can also be expressed as an enrichment factor (, in ‰), which is calculated by

 ) 1000(R - 1)

(3)

Results and Discussion VC Degradation and Isotopic Profiles. The isotopic values of VC plotted versus fraction of VC remaining for all four microbial strains are shown in Figure 2. Initial VC isotope values (δ13C0) for all bottles were within the error of the isotopically characterized laboratory VC standard. In all experiments, the concentration of VC decreased and the isotopic values of the remaining VC became more enriched by 16.1 to 21.6‰ with respect to δ13C0, verifying that aerobic VC degradation is a fractionating process. All media and killed control concentrations remained within 7% of the initial concentration and (0.5‰ of the initial isotope value (Table 2). For each strain, the data fit a Rayleigh model with coefficient of determination for the linear regression (R2) values > 0.98 (Table 1). Values of  for each degradation experiment were calculated using data from all of the replicate culture bottles for each microorganism. Experimentally, R 1 is evaluated as the slope of the line from a plot of ln(R/R0) versus ln f (41), and  is calculated from eq 3. Figure 3 shows an example of this for one of the microorganisms (JS61). Table 1 shows the values of  for each strain. Enrichment Factors for Aerobic VC-Degrading Microorganisms. Enrichment factors for two of the Mycobacterium strains (JS61 and JS617) were the same within 95% confidence intervals (-7.1 ( 0.2 and -7.0 ( 0.3‰, respectively). The experiment with a 4 times larger biomass concentration for JS617 also yielded an identical enrichment factor within 95% confidence intervals (-7.1 ( 0.4‰), suggesting that the extent of fractionation is not a function of the biomass concentration. Similar results were obtained by Mancini et al. (42) for methanogenic biodegradation of benzene. Despite reported differences in growth rates between the Mycobacterium and Nocardioides strains (18), no significant difference was found in the enrichment factor for the Nocardioides strain (-7.6 ( 0.1‰) as compared to the above experiments. Only one isolate, JS60, showed a slightly different enrichment factor (-8.2 ( 0.1‰), but even this is not substantially different from the enrichment factors for the other three microbial strains. In laboratory and field experiments, enrichment factors for anaerobic biodegradation of chlorinated ethenes by different microbial consortia have been shown to have a range for each compound. For example, reported enrichment factors for microbial reductive dechlorination of chlorinated ethenes range from -2.5 to -13.8‰ for TCE (25, 26, 29),

FIGURE 2. Correlation of VC isotopic measurements (symbols) to the Rayleigh model (solid line) for JS60 (A), JS61 (B), JS617 (C), and JS614 (D). Each of the four plots contains data from all replicate bottles for that strain. Vertical error bars represent (0.5 ‰ reproducibility and accuracy for carbon isotopic analysis. Horizontal error bars represent (7% on the fraction of VC remaining. Dashed lines represent (0.5 ‰ error on δ13C0.

TABLE 2. Initial and Final VC Concentrations and Isotope Values for Media and Killed Controls Measured during the Course of Each Microbial Degradationa control type (microbial degradation experiment)b

initial VC concn (µmol/ bottle)

final VC concn (µmol/ bottle)

initial δ13C (‰)

final δ13C (‰)

media (JS60) killed (JS61) media (JS617) killed (JS614)

143 184 163 162

142 192 166 159

-27.9 -28.3 -28.4 -28.4

-28.2 -28.3 -28.4 -28.2

a All controls remained within 7% of initial concentration and (0.5 ‰ of the initial isotope value. b For details of media versus killed control, see text.

from -14.1 to -20.4‰ for cis-DCE (5, 25, 26), and from -21.5 to -31.1‰ for VC (5, 25, 26). Sherwood Lollar et al. (24) used δ13C TCE values from Dover Air Force Base and the range of published values for fractionation of TCE during reductive dechlorination (-2.5 to -13.8‰) to show that the calculated percent degradation for this range of  values resulted in estimates between 40.5% and 66.2%. Similarly, Morrill et al. (21) demonstrated that if the entire range of fractionation factors for cis-DCE (-14.1 to -20.4‰) is used, the total difference in calculated percent degradation is only 37-48%. Compared to this, the range of reported enrichment factors in this study (-7.0 to -8.2‰) is indeed narrow and thus can provide a very robust estimate of the enrichment factor applicable for aerobic VC degradation. Recent work by Chu et al. (43) on stable carbon isotopic fractionation during

FIGURE 3. Calculation of fractionation factor from experimental isotopic data for JS61. The slope of the linear regression through the data is r - 1 and E is calculated from eq 3. The R2 value of 0.995 indicates the data have a good fit to the Rayleigh model. R/R0 is defined in eq 2. aerobic biodegradation of TCE, cis-DCE, and VC supports this as well. Similar to Mycobacterium strains, the Nocardioides strain forms epoxyethane in the first step of its ETH (and chlorooxirane from VC) assimilation pathway (18). However, Coleman et al. (18) also reported that the Nocardioides strain JS614 had a larger maximum specific growth rate (0.71 d-1) and growth yield (10.3 g protein/mol VC) than the other three Mycobacterium strains (0.21-0.23 d-1 and 6.0-6.6 g VOL. 39, NO. 4, 2005 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

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protein/mol VC, respectively) used in the present study, suggesting that the VC assimilating pathway is possibly different between the two types of microorganisms. The isotope data do not provide support for this suggestion of Coleman et al., as there is no substantial difference in the enrichment factors for all four microbial strains. With respect to the initial step in the degradation pathway (i.e., oxidation of VC to chlorooxirane), both Mycobacterium and Nocardioides strains may assimilate VC using the same pathway. Comparison of ∈ between Aerobic and Anaerobic VCDegrading Microorganisms. From a bioremediation perspective, the differences in the enrichment factors among aerobic VC-degrading microbial strains documented in this study are not significant. These enrichment factors are, however, significantly different from those determined for anaerobic VC degradation via reductive dechlorination (range -21.5 to -31.1‰; 5, 25, 26). It is therefore of great interest to determine whether these differences between aerobic and anaerobic enrichment factors can be attributed to different VC degradation pathways, namely, oxidation by a monooxygenase reaction versus reductive dechlorination. Aerobic VCdegrading microorganisms (Mycobacterium and Nocardioides) utilize VC as a carbon source and an electron donor for growth. Coleman et al. (18) report that ETH-grown isolates can be subsequently grown on VC and vice versa without an acclimation period, suggesting the same enzyme (AkMO) is involved in assimilation of both of these compounds. In the aerobic pathway, VC epoxidation is mediated by an alkene monooxygenase enzyme, which breaks a CdC bond to form an epoxide. Conversely, anaerobic reductively dechlorinating pure and mixed cultures use VC as an electron acceptor, resulting in the cleavage of a C-Cl bond. As different bonds are broken, the fractionating step in aerobic and anaerobic VC degradation is different, so that substantially different  values might be expected. Kinetic Isotope Effects. In the case of VC, measured enrichment factors can be directly converted into kinetic isotope effects (KIEs), 12k/13k, according to the formula: 12

k/13k ) 1/(1 + /1000)

(4)

where 12k/13k is the average KIE of both carbon atoms for changes in the bonding at the double bond and  is the measured enrichment factor for a particular process (36). KIEs result from the difference in reaction rates of bonds containing lighter isotopes versus bonds with heavier isotopes, and they therefore depend strongly on the type of bond being broken. The magnitude of KIEs is dependent on the differences in bonding between the transition state and the ground state of the molecule undergoing a particular reaction. In the above equation, secondary isotope effects are assumed to be negligible. Secondary isotope effects involve isotopes not involved in the bond-breaking reaction, as opposed to the type of primary isotope effect described above. Comparison of KIEs of Aerobic and Anaerobic VCDegrading Microorganisms. The KIE for anaerobic VC degradation was calculated using eq 4 to be 1.03 ( 0.007 (based on the range of enrichment factors reported in the literature; 5, 25, 26). From the perspective of chemical bond breakage, the smaller KIE associated with aerobic VC oxidation to chlorooxirane in this study (1.01 ( 0.001) as compared to anaerobic VC reductive dechlorination may agree with KIE theory. Because reaction rate constants are dependent on their activation energies, the KIE can be expressed as (36):

aerobic:

12

k/ k ∝ exp((Ea(13Cd12C) - Ea(12Cd12C))/RT) (5a)

anaerobic:

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12

k/13k ∝ exp((Ea(13C-35Cl) - Ea(12C-35Cl))/RT) (5b)

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where Ea is the activation energy of the molecule (12Cd12C, 13 Cd12C, 12C-Cl, or 13C-Cl) (J mol-1), R is the gas constant (8.314 J mol-1 K-1), and T is the temperature (K). The activation energies are dominated by the energies of molecular movements, of which vibrational dominates over translational and rotational (36). Therefore, the difference (i.e., Ea(13Cd12C) - Ea(12Cd12C) and Ea(13C-35Cl) - Ea(12C-35Cl)) between the two activation energies of isotopic molecules depends strongly on the changes in vibrational frequencies of the bonds involved and the differences between their reduced masses (m1m2/(m1 + m2) where m1 and m2 are the masses of each atom in the bond in question). The more pronounced these changes in vibrational energies, and the heavier the atoms involved in the bond reactions, the more pronounced the isotope effect. The larger fractionation factor observed for reductive dechlorination versus oxidation suggests that, as compared to aerobic oxidation (CdC epoxidation), anaerobic reductive dechlorination (C-Cl cleavage) involves larger differences between the molecules’ reduced masses and larger changes in vibrational frequencies of the bonds involved in the reaction. The anaerobic process involves a larger difference in the reduced mass of the bonds broken because the chlorine atom in 13C-35Cl and 12C-35Cl has a greater mass as compared to the carbon atom in 13Cd12C and 12Cd12C. Further, the overall changes in vibrational frequencies are larger for the anaerobic process. Strictly, the differences in vibrational energies should be calculated as the differences in vibrational energies between the ground state and the transition state. However, because this information is not readily available, we used the vibrational energies of the first intermediates (epoxide for aerobic degradation and assumed carbon radical, C•, for anaerobic degradation) as an estimate for the vibrational energy of the transition state. For C-Cl cleavage to C•, the vibrational frequencies would be 750 cm-1 (typical wavenumber for a C-Cl bond in the ground state (44)) minus 0 cm-1 (bond is completely broken in infinitesimally late transition state), which results in a total difference of 750 cm-1. For CdC epoxidation, the vibrational frequencies may be approximated using stretching frequencies for ethene and epoxyethane (approximately 1620 and 1270 cm-1, respectively (44)), which leads to a much smaller difference of about 350 cm-1. Therefore, for aerobic VC degradation, smaller differences between the reduced masses of the isotopic species and smaller changes in vibrational frequencies of the bonds involved in the reaction tend toward smaller KIEs. The smaller measured enrichment factors found in this study of aerobic VC biodegradation as compared to anaerobic VC biodegradation are consistent with this model. Recently, there has been an increased interest in using laboratory-derived fractionation factors to evaluate biodegradation in the field and to quantitatively estimate the extent of biodegradation using the Rayleigh model (5, 21, 23-26, 28, 29). As noted, the range of anaerobic VC enrichment factors with different microbial consortia (5, 25, 26) is relatively small. The aerobic enrichment factor range determined in this study is even smaller, suggesting that the calculated enrichment factor is independent of the particular VC-assimilating strain present at the site. Finally, significant fractionation occurs during aerobic VC biodegradation, suggesting that stable carbon isotope measurements can be used to distinguish between biodegraded and nonbiodegraded VC in aerobic and anaerobic environments.

Acknowledgments This work was funded by the National Science and Engineering Research Council Strategic Projects Grant. Participation of the Cornell researchers was made possible by funding from the U.S. Strategic Environmental Research and Development Program.

Literature Cited (1) Squillace, P. J.; Moran, M. J.; Lapham, W. W.; Price, C. V.; Clawges, R. M.; Zogorski, J. S. Volatile organic compounds in untreated ambient groundwater of the United States, 1985-1995. Environ. Sci. Technol. 1999, 33, 4176-4187. (2) Kielhorn, J.; Melber, C.; Wahnschaffe, U.; Aitio, A.; Mangelsdorf, I. Vinyl chloride: Still a cause for concern. Environ. Health Perspect. 2000, 108, 579-607. (3) He, J.; Ritalhti, K. M.; Yang, K.-L.; Koenigsberg, S. S.; Loffler, F. E. Detoxification of vinyl chloride to ethene coupled to growth of an anaerobic bacterium. Nature 2003, 424, 62-65. (4) Maymo-Gatell, X.; Chein, Y.-t.; Gossett, J. M.; Zinder, S. H. Isolation of a bacterium that reductively dechlorinates tetrachloroethene to ethene. Science 1997, 276, 1568-1571. (5) Hunkeler, D.; Aravena, R.; Cox, E. Carbon isotopes as a tool to evaluate the origin and fate of vinyl chloride: Laboratory experiments and modeling of isotope evolution. Environ. Sci. Technol. 2002, 36, 3378-3384. (6) Lorah, M. M.; Olsen, L. D. Degradation of 1,1,2,2-tetrachloroethane in a freshwater tidal wetland: Field and laboratory evidence. Environ. Sci. Technol. 1999, 33, 227-234. (7) Bradley, P. M.; Chappelle, F. H. Acetogenic microbial degradation of vinyl chloride. Environ. Sci. Technol. 2000, 34, 2761-2763. (8) Ellis, D. E.; Lutz, E. J.; Odom, J. M.; Buchanan, R. J., Jr.; Bartlett, C. L.; Lee, M. D.; Harkness, M. R.; Deweerd, K. A. Bioaugmentation for accelerated in situ anaerobic bioremediation. Environ. Sci. Technol. 2000, 34, 2254-2260. (9) Major, D. W.; McMaster, M. L.; Cox, E. E.; Edwards, E. A.; Dworatzek, S. M.; Hendrickson, E. R.; Starr, M. G.; Payne, J. A.; Buonamici, L. W. Field demonstration of successful bioaugmentation to achieve dechlorination of tetrachloroethene to ethene. Environ. Sci. Technol. 2002, 36, 5106-5116. (10) Hendrickson, E. R.; Payne, J. A.; Young, R. M.; Starr, M. G.; Perry, M. P.; Fahnestock, S.; Ellis, D. E.; Ebersole, R. C. Molecular analysis of Dehalococcoides 16S ribosomal DNA from chloroethene-contaminated sites throughout North America and Europe. Appl. Environ. Microbiol. 2002, 68, 485-495. (11) Harkness, M. R.; Bracco, A. A.; Brennan, M. J.; Deweerd, K. A.; Spivack, J. L. Use of bioaugmentation to stimulate complete reductive dechlorination of trichloroethene in Dover soil columns. Environ. Sci. Technol. 1999, 33, 1100-1109. (12) Shim, H.; Ryoo, D.; Barbieri, P.; Wood, T. K. Aerobic degradation of mixtures of tetrachloroethylene, trichloroethylene, dichloroethylenes, and vinyl chloride by toluene-o-xylene monooxygenase of Pseudomonas stutzeri OX1. Appl. Microbiol. Biotechnol. 2001, 56, 265-269. (13) Verce, M. F.; Freedman, D. L. Modeling the kinetics of vinyl chloride cometabolism by an ethane-grown Pseudomonas sp. Biotechnol. Bioeng. 2001, 71, 274-285. (14) Hartmans, S.; de Bont, J. A. M.; Tramper, J.; Luyben, K. Ch. A. M. Bacterial degradation of vinyl chloride. Biotechnol. Lett. 1985, 7, 383-388. (15) Hartmans, S.; de Bont, J. A. M. Aerobic vinyl chloride metabolism in Mycobacterium aurum L1. Appl. Environ. Microbiol. 1992, 58, 1220-1226. (16) Verce, M. F.; Ulrich, R. L.; Freedman, D. L. Characterization of an isolate that uses vinyl chloride as a growth substrate under aerobic conditions. Appl. Environ. Microbiol. 2000, 66, 35353542. (17) Verce, M. F.; Ulrich, R. L.; Freedman, D. L. Transition from cometabolic to growth-linked biodegradation of vinyl chloride by a Pseudomonas sp. isolated on ethene. Environ. Sci. Technol. 2001, 35, 4242-4251. (18) Coleman, N. V.; Mattes, T. E.; Gossett, J. M.; Spain, J. C. Phylogenetic and kinetic diversity of aerobic vinyl chlorideassimilating bacteria from contaminated sites. Appl. Environ. Microbiol. 2002, 68, 6162-6171. (19) Coleman, N. V.; Spain, J. C. Epoxyalkane: Coenzyme M transferase in the ethene and vinyl chloride biodegradation pathways of Mycobacterium strain JS60. J. Bacteriol. 2003, 185, 5536-5545. (20) Coleman, N. V.; Spain, J. C. Distribution of the Coenzyme M pathway of epoxide metabolism among ethene- and vinyl chloride-degrading Mycobacterium strains. Appl. Environ. Microbiol. 2003, 69, 6041-6046. (21) Morrill, P.; Lacrampe-Couloume, G.; Slater, G. F.; Sleep, B.; Edwards, E. A.; McMaster, M.; Major, D. W.; Sherwood Lollar, B. Quantifying chloroethene mass degraded during reductive dechlorination using stable carbon isotopes at Kelly AFB: Comparison to concentration-derived estimates. J. Contam. Hydrol. (in press).

(22) Hirschorn, S. K.; Dinglasan, M. J.; Elsner, M.; Mancini, S. A.; Lacrampe-Couloume, G.; Edwards, E. A.; Sherwood Lollar, B. Pathway dependent isotopic fractionation during aerobic biodegradation of 1,2-dichloroethane. Environ. Sci. Technol. 2004, 38, 4775-4781. (23) Song, D. L.; Conrad, M. E.; Sorenson, K. S.; Alvarez-Cohen, L. Stable carbon isotope fractionation during enhanced in situ bioremediation of trichloroethene. Environ. Sci. Technol. 2002, 36, 2262-2268. (24) Sherwood Lollar, B.; Slater, G. F.; Sleep, B.; Witt, M.; Klecka, G. M.; Harkness, M.; Spivack, J. Stable carbon isotope evidence for intrinsic bioremediation of tetrachloroethene and trichloroethene at Area 6, Dover Air Force Base. Environ. Sci. Technol. 2001, 35, 261-269. (25) Slater, G. F.; Sherwood Lollar, B.; Sleep, B. E.; Edwards, E. A. Variability in carbon isotopic fractionation during biodegradation of chlorinated ethenes: Implications for field applications. Environ. Sci. Technol. 2001, 35, 901-907. (26) Bloom, Y.; Aravena, R.; Hunkeler, D.; Edwards, E.; Frape, S. K. Carbon isotope fractionation during microbial dechlorination of trichloroethene, cis-1,2-dichloroethene, and vinyl chloride: Implications for assessment of natural attenuation. Environ. Sci. Technol. 2000, 34, 2768-2772. (27) Hunkeler, D.; Aravena, R. Evidence of substantial carbon isotope fractionation among substrate, inorganic carbon, and biomass during aerobic mineralization of 1,2-dichloroethane by Xanthobacter autotrophicus. Appl. Environ. Microbiol. 2000, 66, 4870-4876. (28) Hunkeler, D.; Aravena, R.; Butler, B. J. Monitoring microbial dechlorination of tetrachloroethene (PCE) in groundwater using compound-specific stable carbon isotope ratios: Microcosm and field studies. Environ. Sci. Technol. 1999, 33, 2733-2738. (29) Sherwood Lollar, B.; Slater, G. F.; Ahad, J.; Sleep, B.; Spivack, J.; Brennan, M.; MacKenzie, P. Contrasting carbon isotope fractionation during biodegradation of trichloroethylene and toluene: Implications for intrinsic bioremediation. Org. Geochem. 1999, 30, 813-820. (30) Schuth, C.; Taubald, H.; Bolano, N.; Maciejczyk, K. Carbon and hydrogen isotope effects during sorption of organic contaminants on carbonaceous materials. J. Contam. Hydrol. 2003, 64, 269-281. (31) Huang, L.; Sturchio, N. C.; Abrajano, T., Jr.; Heraty, L. J.; Holt, B. D. Carbon and chlorine isotope fractionation of chlorinated aliphatic hydrocarbons by evaporation. Org. Geochem. 1999, 30, 777-785. (32) Slater, G. F.; Ahad, J. M. E.; Sherwood Lollar, B.; Allen-King, R.; Sleep, B. Carbon isotope effects resulting from equilibrium sorption of dissolved VOCs. Anal. Chem. 2000, 72, 5669-5672. (33) Slater, G. F.; Dempster, H. S.; Sherwood Lollar, B.; Ahad, J. Headspace analysis: A new application for isotopic characterization of dissolved organic contaminants. Environ. Sci. Technol. 1999, 33, 190-194. (34) Poulson, S. R.; Drever, J. I. Stable isotope (C, Cl and H) fractionation during vaporization of trichloroethylene. Environ. Sci. Technol. 1999, 33, 3689-3694. (35) Galimov, E. M. The Biological Fractionation of Isotopes; Academic Press: Orlando, FL, 1985. (36) Melander, L.; Saunders, W. H., Jr. Reaction Rates of Isotopic Molecules; John Wiley and Sons: New York, 1980. (37) Madigan, M. T.; Martinko, J. M.; Parker, J. Brock Biology of Microorganisms, 9th ed.; Prentice Hall: Upper Saddle River, NJ, 2000. (38) Bradford, M. A rapid and sensitive method for the quantitation of microgram quantities of protein utilizing the principle of protein-dye binding. Anal. Biochem. 1976, 72, 248-254. (39) Dempster, H. S.; Sherwood Lollar, B.; Feenstra, S. Tracing organic contaminants in groundwater: A new methodology using compound specific isotopic analysis. Environ. Sci. Technol. 1997, 31, 3193-3197. (40) Mancini, S. A.; Lacrampe-Couloume, G.; Jonker, H.; van Breukelen, B. M.; Groen, J.; Volkering, F.; Sherwood Lollar, B. Hydrogen isotopic enrichment: An indicator of biodegradation at a petroleum hydrocarbon contaminated field site. Environ. Sci. Technol. 2002, 36, 2464-2470. (41) Mariotti, A.; Germon, J. C.; Hubert, P.; Kaiser, P.; Letolle, R.; Tardieux, A.; Tardieux, P. Experimental determination of nitrogen kinetic isotope fractionation: Some principles; VOL. 39, NO. 4, 2005 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

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Illustration for the denitrification and nitrification processes. Plant Soil 1981, 62, 413-430. (42) Mancini, S. A.; Ulrich, A. C.; Lacrampe-Couloume, G.; Sleep, B. E.; Edwards, E. A.; Sherwood Lollar, B. Carbon and hydrogen isotope fractionation during anaerobic biodegradation of benzene. Appl. Environ. Microbiol. 2003, 69, 191-198. (43) Chu, K.-H.; Mahendra, S.; Song, D. L.; Conrad, M. E.; AlvarezCohen, L. Stable carbon isotope fractionation during aerobic biodegradation of chlorinated ethenes. Environ. Sci. Technol. 2004, 38, 3126-3130.

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(44) Pavia, D. L.; Lampman, G. M.; Kriz, G. S. Introduction to Spectroscopy, 2nd ed.; Saunders College Publishing: Orlando, FL, 1996.

Received for review May 11, 2004. Revised manuscript received November 2, 2004. Accepted November 4, 2004. ES0492945