Environ. Sci. Technol. 2001, 35, 901-907
Variability in Carbon Isotopic Fractionation during Biodegradation of Chlorinated Ethenes: Implications for Field Applications GREGORY F. SLATER,† B A R B A R A S H E R W O O D L O L L A R , * ,† BRENT E. SLEEP,‡ AND ELIZABETH A. EDWARDS§ Stable Isotope Laboratory, Department of Geology, University of Toronto, 22 Russell Street, Toronto, Canada M5S 3B1, Department of Civil Engineering, University of Toronto, Toronto, Ontario, Canada M5S 1A4, and Department of Chemical Engineering & Applied Chemistry, University of Toronto, Toronto, Ontario, Canada, M5S 3E5
Stable carbon isotopic analysis has the potential to assess biodegradation of chlorinated ethenes. Significant isotopic shifts, which can be described by Rayleigh enrichment factors, have been observed for the biodegradation of trichloroethlyene (TCE), cis-dichloroethylene (cDCE), and vinyl chloride (VC). However, until this time, no systematic investigation of isotopic fractionation during perchloroethylene (PCE) degradation has been undertaken. In addition, there has been no comparison of isotopic fractionation by different microbial consortia, nor has there been a comparison of isotopic fractionation by consortia generated from the same source, but growing under different conditions. This study characterized carbon isotopic fractionation during reductive dechlorination of the chlorinated ethenes, PCE in particular, for microbial consortia from two different sources growing under different environmental conditions in order to assess the extent to which different microbial consortia result in different fractionation factors. Rayleigh enrichment factors of -13.8‰, -20.4‰, and -22.4‰ were observed for TCE, cDCE, and VC, respectively, for dechlorination by the KB-1 consortium. In contrast, isotopic fractionation during reductive dechlorination of perchloroethylene (PCE) could not always be approximated by a Rayleigh model. Dechlorination by one consortium followed Rayleigh behavior ( ) -5.2), while a systematic change in the enrichment factor was observed over the course of PCE degradation by two other consortia. Comparison of all reported enrichment factors for reductive dechlorination of the chlorinated ethenes shows significant variation between experiments. Despite this variability, these results demonstrate that carbon isotopic analysis can provide qualitative evidence of the occurrence and relative extent of microbial reductive dechlorination of the chlorinated ethenes.
Introduction Monitoring contaminant degradation is an important part of successful groundwater remediation strategies. A key issue in quantification of degradation is distinguishing between contaminant attenuation due to nondegradative processes 10.1021/es001583f CCC: $20.00 Published on Web 02/02/2001
2001 American Chemical Society
such as volatilization, sorption and dilution, and attenuation due to biotic degradation processes (1). Natural attenuation, or intrinsic remediation, involves the reduction of dissolved concentrations of a contaminant by a combination of naturally occurring degradative and nondegradative processes (2). Though both of these types of processes will attenuate contaminant movement in the subsurface, degradation is the only process by which a contaminant is transformed, ideally to nontoxic end products. Degradation of chlorinated ethenes can occur abiotically via inorganic catalysis; however, this process is much slower in natural environments than in engineered systems such as iron wall remediation schemes (3). In contrast, anaerobic microbes have been isolated which are capable of complete and/or partial dechlorination of chlorinated ethenes at relatively high rates (e.g. refs 4-6). Quantifying intrinsic biodegradation presently relies on accumulation of a broad basis of evidence ranging from microbiological and chemical field measurements, laboratory and field experiments, and modeling (2). Current difficulties in this approach include obtaining accurate mass balances of contaminants, electron donors, and end products in heterogeneous aquifer systems, distinguishing between biodegradation effects and decreases in contaminant concentrations due to physical processes such as sorption, volatilization and dilution, and difficulty in extrapolating laboratory-based microbiological studies to field situations (1, 7). Stable carbon isotopic analysis of dissolved organic contaminants has the potential to directly assess the extent of intrinsic bioremediation of chlorinated ethenes occurring in the subsurface (8-10). Stable carbon isotopic analysis involves measurement of the ratio of the two stable isotopes of carbon present in a sample. This ratio is expressed as a δ13C value where
δ13C ) (Rsample/Rstandard - 1) × 1000
(1)
Rsample is the 13C/12C ratio in a given sample, and Rstandard is the 13C/12C ratio in a standard reference material, in this case V-PDB. For the chlorinated ethenes and BTEX compounds, nondegradative subsurface processes such as dissolution, volatilization from aqueous solution, and adsorption do not involve isotopic fractionation greater than 0.5‰ (the typical accuracy and reproducibility of continuous flow isotope analysis techniques) (11-13). In contrast, recent studies have shown that microbial reductive dechlorination of chlorinated ethenes and other chlorinated compounds causes fractionation of the isotopic composition of the residual dissolved chlorinated contaminant (8-10). This suggests that changes in isotopic values of chlorinated solvents during degradation can be used to identify the occurrence of degradation independent of nonfractionating, nondegradative processes of contaminant attenuation. Isotopic fractionation during biodegradation is a result of preferential degradation of molecules containing the lighter isotope (12C). The presence of the heavier carbon 13 isotope results in a higher bond vibrational frequency which requires higher activation energy to react (14). Significantly, the relationship between isotopic fractionation and extent of * Corresponding author phone: (416)978-0770; fax: (416)978-3938; e-mail:
[email protected]. † Stable Isotope Laboratory, Department of Geology. ‡ Department of Civil Engineering. § Department of Chemical Engineering & Applied Chemistry. VOL. 35, NO. 5, 2001 / ENVIRONMENTAL SCIENCE & TECHNOLOGY
9
901
reductive dechlorination has been modeled by a simple isotopic model, known as the Rayleigh model, for both abiotic and biotic reductive dechlorination of TCE (8, 9, 15, 16). The Rayleigh model assumes a constant isotopic preference during a reaction, which is reflected in a fractionation factor, R. The R factor relates the ratio of isotopic composition of the substrate at a given time, R, to the initial isotopic composition, Ro, to the fraction of substrate remaining (f) where
R/Ro ) f(R-1)
(2)
The isotopic preference of a reaction can also be expressed as an enrichment factor, , where ) 1000(R-1). At this time, enrichment factors for reductive dechlorination have only been derived for a few microbial enrichment cultures. For dechlorination of TCE, Sherwood Lollar et al. (8) determined a reproducible enrichment factor using a sulfate-reducing culture enriched for growth on TCE with lactate as an electron donor. Bloom et al. (9) observed variation in enrichment factors for replicate experiments degrading TCE, cDCE, and VC with the methanogenic consortium KB-1. Hunkeler et al. (10) used a methanogenic culture growing on PCE with methanol as the electron donor and estimated an enrichment factor for dechlorination of PCE, TCE, cDCE, and VC. However, these estimates were for compounds undergoing both production and consumption simultaneously and are not accurate Rayleigh enrichment factors. The difference in enrichment factors observed by Sherwood Lollar et al. (8) and Bloom et al. (9) suggest that differences in enrichment factors do occur between microbial consortia and even for replicate degradations by the same microbial consortia. Understanding the extent to which this variation occurs and why it is essential to the application of stable isotopic analysis as a tool to investigate and monitor biodegradation of chlorinated solvents in the field. The objective of this study was to assess the extent of variation in enrichment factors for anaerobic reductive dechlorination of the chlorinated ethenes. The first objective was to assess the isotopic fractionation of all chlorinated ethenes (PCE, TCE, cDCE, and VC) by the methanogenic KB-1 consortium. The second objective was to investigate the influence of electron donor on isotopic fractionation using two consortia generated from the same source, growing in the same media, but utilizing different electron donors. Finally, a comparison was made of all reported enrichment factors for anaerobic dechlorination of the chlorinated ethenes in order to estimate the variability which might be observed in the field.
Experimental Approach Two experiments were carried out for this study. Experiment 1 characterized the carbon isotopic fractionation of PCE, TCE, cDCE, and VC during reductive dechlorination by the KB-1 enriched microbial consortium which has been subcultured to grow on each of the chlorinated ethenes. Experiment 2 evaluated isotopic fractionation during reductive dechlorination of PCE by the TP microbial consortium that was enriched using different electron donors. Experiment 1: Comparison of Fractionation of PCE, TCE, cDCE, and VC. The KB-1 consortium was generated from a TCE-contaminated field site in Southern Ontario. Rapid dechlorination was observed within 2 weeks of initial establishment of microcosms with site groundwater and soil, amended with a mixture of electron donors (methanol, ethanol, lactate, and acetate) (17). Subcultures were developed using methanol as the electron donor and one of either PCE, TCE, cDCE, or VC as electron acceptor. These subcultures have been maintained in defined mineral medium (18) for over 2 years. 902
9
ENVIRONMENTAL SCIENCE & TECHNOLOGY / VOL. 35, NO. 5, 2001
The isotopic effects of anaerobic reductive dechlorination of PCE, TCE, cDCE, and VC by these subcultures of the same microbial consortium were characterized. For this experiment each subculture was split into two replicates, A and B. These samples consisted of 100 mL of culture transferred to an autoclaved 250 mL serum bottle and capped with a screw cap Mininert valve (Precision Sampling Corp.). Prior to addition of chlorinated ethenes, the cultures were purged with N2 for 20 min to remove any residual chlorinated ethenes from previous experiments. The chlorinated ethenes and methanol (electron donor) were then added to the experimental and control bottles. For PCE, TCE, and cDCE 20 µL of a 1:9 (v/v) stock solution of each chlorinated ethene in methanol was added to respective bottles. For the bottles amended with VC, 750 µL of VC (gas) and 30 µL of methanol (neat) were added. These additions corresponded to aqueous concentrations of 94, 140, 212, 115 µM or 16, 18, 21, and 7.2 mg/L, respectively, using dimensionless Henry’s law constants of 0.723, 0.392, 0.167, and 1.137 (19) and assuming dissolution in deionized water. The concentration of the methanol electron donor was 4.45 mmol/L. Two control bottles containing the subculture poisoned with 100 µL of a 5% solution of HgCl2 and amended with chlorinated ethene stock solutions and two abiotic control bottles containing only deionized water and the chlorinated ethene stock solution were also prepared. The cultures were incubated inverted at 22 ( 1 °C on an orbital shaker at 150 rpm in a fume hood. Each replicate bottle (A,B) was run twice for a total of four repetitions (A1, A2, B1, B2) of each degradation experiment carried out with each chlorinated ethene. Experimental and control bottles were sampled repeatedly over the duration of the experiment to determine both the concentration and isotopic composition of the chlorinated ethenes. Experiment 2: Influence of Electron Donor on Fractionation of PCE. The influence of electron donor on isotopic fractionation of PCE was compared using consortia enriched from samples of biomass obtained from an anaerobic digester at the Metropolitan Toronto Main Treatment Plant. These consortia were enriched in anaerobic basal medium (20) with PCE and ethanol (TP-ethanol) or butryic acid (TP-butyric acid). The TP-butyric acid consortium was originally enriched on PCE and methanol as the electron donor (21) and subsequently switched to butyric acid. The TP-ethanol and TP-butyric acid consortia were split into replicates (A,B) consisting of 175 mL of culture in an autoclaved 250 mL serum bottle capped with a screw cap Mininert valve (Precision Sampling Corp.). The cultures were fed with 2 µL of neat PCE and 20 µL of neat ethanol or butyric acid. These additions corresponded to an aqueous concentration of 95 µM (16 mg/L) and an initial electron donor concentration of 3.4 mM ethanol or 2.2 mM butyric acid. Two control bottles containing the culture poisoned with 100 µL of 5% HgCl2 solution amended with PCE and electron donor and two abiotic controls containing deionized water, PCE, and electron donor were also prepared. Experimental bottles were again incubated inverted at 22 ( 1 °C on an orbital shaker at 150 rpm in a fume hood. Experimental and control bottles were sampled repeatedly over the duration of reaction to determine both the concentration and isotopic composition of the chlorinated ethenes. A total of two repetitions of PCE degradation was carried out using each of the TP consortia. Analytical Methods. The total moles of chlorinated ethene in each bottle was analyzed by removing 300 µL of headspace from each bottle using a 500 µL Pressure-Lok gastight syringe (Vici Precision Sampling Inc.) and injecting the sample into a Varian 3300 GC equipped with a 0.53 mm × 30 m GS-Q column (J&W Scientific) and an FID detector. The GC oven temperature program was 70 °C hold 1 min, increase to 200
°C at 26°/min, then increase to 225 °C at 5°/min, and hold 2 min. Reproducibility on standard analyses was ( 5%. Isotopic analysis was carried out by headspace analysis, as described in Slater et al. (11) who demonstrated no significant isotopic fractionation (>0.5‰) for volatilization of TCE from aqueous solution. Further laboratory tests obtained the same results for the other chlorinated ethenes (unpublished data). In this study, headspace analysis consisted of removing 300-1000 µL of headspace from each bottle using a 1000 µL Pressure-Lok gastight syringe (Vici Precision Sampling Inc.) and injecting into a Gas Chromatograph-Combustion-Isotope Ratio Mass Spectrometer (GC-C-IRMS) system. The GC-C-IRMS system consisted of a Varian 3400 Gas Chromatograph equipped with a capillary column coupled via a combustion interface to a Finnigan MAT 252 gas source isotope ratio mass spectrometer. For isotopic analysis of PCE, TCE, cDCE, and VC a 30 m × 0.25 mm DB-624 column (J&W Scientific) with a 1.4 µm film thickness was used. The temperature program for the DB-624 column was -15 °C hold 5 min, increase to 90 C at 10°/min, and hold 5 min. For isotopic analysis of ethene a 30 m × 0.25 mm Supel-Q plot column with a 10 µm film thickness (Supelco Inc.) was used. The temperature program for the Supel-Q column was 40 °C hold 4 min, increase at 15°/min to 150 °C, and hold 5 min. Samples were run against external CO2 isotopic standards, and all δ13C values are reported relative to the V-PDB standard. While internal reproducibility of based on duplicate injections of a given sample is 0.1-0.3‰, differences between samples (error bars) are assigned a value of 0.5‰ to incorporate both reproducibility and accuracy associated with GC-C-IRMS analysis after Dempster et al. (22) and Slater et al. (11). In particular, changing signal sizes over the course of degradation can result in small variations introduced into GC-C-IRMS analysis. This is supported by the findings of Hunkeler et al. (10) who observed a 0.3‰ variation in isotopic composition of standards over an order of magnitude variation in signal size. Injection volumes were varied to minimize the impact of signal size on isotopic composition. Total mass removed during sampling was calculated to be in the nanomole range and did not have significant impact on mass or isotopic balances.
Results Concentration Profiles. Figures 1 and 2 show total µmol per bottle and isotopic profiles for the reductive dechlorination of PCE and TCE, respectively, by the KB-1 consortium. Dechlorination of cDCE and VC by KB-1 also resulted in complete conversion to ethene (data not shown). For each chloroethene, the four repetitions carried out yielded essentially the same sequential degradation profile with complete conversion to ethene, so only one data set is plotted. Ethene was not quantified for the PCE experiment (Figure 1a) but was for the TCE (Figure 2a), cDCE, and VC experiments. Similar dechlorination profiles were observed for the PCE-fed TP-ethanol and TP-butyric acid consortia (data not shown). No consistent chloroethene loss ((5%) was observed in the control bottles (Figures 1a and 2a). The sum of moles of chloroethenes plus ethene during dechlorination was within ( 10% of the initial moles of chloroethene added at each data point for the experiments amended with VC and within (20% of the initial moles for all but a few sampling points in the experiments amended with TCE or cDCE. For the TCE subculture the final µmoles of ethene closely approached the initial µmoles of TCE, indicating quantitative conversion (Figure 2a). The mole balance for the PCE subculture could not be assessed since ethene was not quantified (Figure 1a). Zero-order rate constants and correlation factors (r2) are shown in Table 1 for each repetition.
FIGURE 1. a: Total mass (in µmol) of chlorinated ethenes per bottle versus time for reductive dechlorination of PCE by the KB-1 consortium. Error bars represent ( 5% reproducibility for GC analysis. b: δ13C values (‰) for the chlorinated ethenes versus time for reductive dechlorination of PCE by the KB-1 consortium. Error bars represent ( 0.5‰ accuracy and reproducibility for GC-CIRMS analysis and are smaller than the plotted symbols. Throughout the experiment, δ13C for PCE in control bottles remained identical to δ13C of the initial PCE. Isotopic Profiles. Figures 1b and 2b show δ13C values of the chlorinated ethenes plotted versus time during degradation of PCE and TCE, respectively, by the KB-1 consortium (Experiment 1). As for the concentration data, the δ13C profiles for the other experiments (cDCE, VC, TP-butyric acid, TPethanol) followed the same trends seen in these figures and are consequently not shown. Isotopic mass balance was not possible for the PCE experiments as ethene was not quantified. For the other experiments, slight enrichment of the isotopic mass balance toward the end of the experiment was observed but was likely within error (Figure 2b). The variability in the concentration mole balances ((20%) may have affected total isotopic mass balances. However, the reproducibility of the isotopic data for each parent compound ( for most replicates within 95% confidence intervals of each other) demonstrates that there is no significant effect on enrichment factors due to mass balance issues (Table 1). Similarly, the trends of isotopic enrichment over the course of degradation would not be significantly affected. For all of the experiments, (KB-1 PCE, KB-1 TCE, KB-1 cDCE, KB-1 VC, TP-ethanol PCE, and TP-butyric acid PCE) a trend of increasing isotopic enrichment (more positive δ13C values) with progressive biodegradation was observed, both for the initial chlorinated ethene and for all dechlorination intermediates. In addition, at any given sampling point, the isotopic composition of the dechlorination product was isotopically depleted in the heavier isotope with respect to the reactant. These trends are consistent with preferential VOL. 35, NO. 5, 2001 / ENVIRONMENTAL SCIENCE & TECHNOLOGY
9
903
TABLE 1. Enrichment Factors and Rate Constants for All Repetitionsa repetition A1 con- chlorosortium ethene KB-1
PCE
KB-1 KB-1 KB-1 TP TP
TCE cDCE VC PCE PCE
e- donor
rate
methanol
0.11
r2
E
0.93 -5.5
repetition B1 95% CI
rate
(1.4 0.15
r2
repetition A2 95% CI rate
E
0.98 -5.5
r2
E
(1.1 5.5 0.84 -2.6
repetition B2 95% CI
(-12.1, -0.06) methanol 58 0.99 -14.3 (1.5 50 0.99 -13.4 (2.2 12 0.97 -13.9 (0.6 methanol 23 0.99 -21.9 (5.2 21 0.99 -25.5 (1.2 4.6 0.46 -18.8 (0.9 methanol 0.46 0.77 -21 (1.7 na na na na 1.0 0.99 -26.0 (5.5 butyric acid 0.084 0.99 -5.4 (0.6 0.098 0.93 -5.1 (0.2 na na na na ethanol 0.67 0.94 -1.8 (1.1 0.72 0.96 -3.0 (1.5 na na na na
rate
r2
E
95% CI
6.5 0.97 -2.6
(-10.1, -0.05) 21 0.98 -15.2 (6.1 5.0 0.86 -18.9 (1.0 5.3 0.97 -22.2 (2.5 na na na na na na na na
a Enrichment factors ( in ‰) and zero-order rate constants (µmol/bottle/day) calculated for reductive dechlorination of the chlorinated ethenes during each experimental repetition for each consortia/donor/acceptor combination. For zero-order rate constants, correlation factors (r2) are shown. Data points used to calculate each are shown in Figure 3. Enrichment factors are calculated by plotting the data on a plot of ln f vs ln(((δ13CTCE/1000)+1)/((δ13Co/1000)+1)) after Mariotti et al. (23) and determining the slope (R-1) and 95% confidence intervals by least-squares regression. Fractionation factors are converted to permil enrichment factors, ) 1000(R-1). In cases where 95% confidence intervals were not symmetrical around data, upper and lower 95% values are shown in parentheses. na indicates no data available. In the case of KB-1 VC (B1) the sample was lost. For the TP consortia only two repetitions were carried out. While Rayleigh model does not agree well with KB-1 PCE or TP-ethanol PCE cultures (see text), are included for these experiments for consistency.
subculture (Figure 2b), where quantitative conversion was achieved, the final isotopic composition of the end-product ethene is the same as the initial isotopic composition of TCE. All control data was within 0.5‰ of initial isotopic values for all experiments (Figures 1b and 2b).
Discussion Significant isotopic fractionation of the residual chlorinated ethenes was observed in all experiments. Comparison of isotopic fractionation during reductive dechlorination by these different microbial consortia can most effectively be achieved by applying the Rayleigh model. The Rayleigh model can be applied to an irreversible reaction where the fractionation factor, R, is characteristic of the isotopic fractionation during that reaction (23). Because R is characteristic of the reaction, and does not depend on the extent to which the reaction proceeds, fractionation factors are an effective means of comparing the isotopic fractionation occurring during different reactions. Fractionation factors were determined for each of the donor/acceptor/culture combinations using the data from all four replicates of each experiment and the approach of Mariotti et al. (23) shown in eq 3
(R-1)ln f ) ln(((δ13CTCE/1000) + 1)/((δ13CTCEo/1000) + 1)) (3)
FIGURE 2. a: Total mass (in µmol) of chlorinated ethenes per bottle versus time for reductive dechlorination of TCE by the KB-1 consortium. Error bars represent ( 5% reproducibility for GC analysis. b: δ13C values for the chlorinated ethenes versus time for reductive dechlorination of TCE by the KB-1 consortium. Error bars represent ( 0.5‰ accuracy and reproducibility for GC-C-IRMS analysis and are smaller than the plotted symbols. Dashed line shows isotopic mass balance. Throughout the experiment, δ13C for TCE in control bottles remained identical to δ13C of the initial TCE. reaction of the isotopically lighter reactant, resulting in enrichment of the reactant pool in the heavier isotope (13C). Only in Figure 1b is the isotopic composition of the TCE initially produced more isotopically enriched than the PCE from which it was formed. This is likely due to the fact that the TCE has already started to undergo dechlorination to cDCE. This behavior was only observed in the PCE degrading consortia and may be because the dechlorination of TCE to cDCE is much more strongly isotopically fractionating than the dechlorination of PCE to TCE. Finally, for the TCE KB-1 904
9
ENVIRONMENTAL SCIENCE & TECHNOLOGY / VOL. 35, NO. 5, 2001
where δ13CTCE and δ13CTCEo are the isotopic compositions of the TCE at a given time and the initial time, respectively, and where f ) CTCE/CTCEo (the fraction of the original TCE remaining). A least-squares regression of a plot of ln f vs ln(((δ13CTCE/1000) + 1)/((δ13CTCEo/1000) +1))yields a slope of R-1. Fractionation factors were converted to permil enrichment factors, ) 1000(R-1), for comparison. The isotopic data for all replicates and all experiments were plotted versus the fraction remaining (f) (Figure 3 a-f). The best fit Rayleigh model curves were computed and plotted on the same figures using the enrichment factors derived from all the data for each experiment as described above. For the TCE, cDCE, and VC experiments (Figure 3b-d), the Rayleigh model agreed very well with the data from all experimental replicates. In contrast, for the PCE datasets good agreement between the Rayleigh model and the data was only obtained for the TP-butyric acid culture (Figure 3f). The Rayleigh model did not agree well with the data for the KB-1 PCE culture (Figure 3a) or the TP-ethanol culture (Figure 3e). For these two experiments, the data follows the Rayleigh curve initially, but a systematic shift occurs at ∼50% fraction remaining for all replicates. In the latter part of these two experiments, the data fell below the Rayleigh model curve.
FIGURE 3. Rayleigh enrichment factors and experimental data. δ13C values of the residual chlorinated ethene versus fraction remaining during reductive dechlorination of (a) PCE; (b) TCE; (c) cDCE; (d) VC by the KB-1 consortium and PCE by the TP-ethanol consortium (e) and the TP-butyric consortium (f). For each compound and consortium, data from all experiments and replicates are plotted. Error bars represent ( 0.5‰ accuracy and reproducibility for GC-C-IRMS analysis. The curves represent Rayleigh models for each culture using enrichment factors determined from all data points (Table 1). Comparison of Fractionation of PCE, TCE, cDCE, and VC. The enrichment factors and 95% confidence intervals for each repetition with PCE, TCE, cDCE, and VC are shown in Table 1. In general, for TCE, cDCE, and VC, enrichment factors were reproducible (identical within 95% confidence intervals) for all repetitions (A1, A2, B1, B2). In the case of cDCE there was some difference in the enrichment factor between repetitions (1 and 2). While the enrichment factor for each repetition varied from -21.9‰ (A1) and -25.5‰ (B1) to -18.8‰ (A2) and -18.9‰ (B2), all four values are within the 95% confidence interval for the value for repetition A1. Due to this reproducibility between enrichment factors for different repetitions, all data from all replicates was combined to determine overall enrichment factors for each donor/acceptor/culture combination for each chlorinated ethene (Table 2). The tight 95% confidence intervals shown in Table 2 for these regression curves support the fact that these combined enrichment factors are descriptive of the isotopic fractionation produced by the consortia. Even for cDCE, where the most variation in enrichment factor was observed between replicates, the 95% confidence interval for the combined enrichment factor is still small ((1.8). Correlation factors (r2) in Table 2 indicate whether the data can be described by a Rayleigh model. Clearly the reductive dechlorination of TCE, cDCE, and VC show Rayleigh fractionation (high r2 values), while dechlorination of PCE does not consistently agree with a Rayleigh model (low r2 values) for KB-1 PCE and TP-ethanol PCE.
TABLE 2. Combined Enrichment Factors for Each Consortia Used in This Study Generated by Mariotti et al. (23) Least-Squares Regression on All Data for Each Consortiaa consortia
e- donor
E
95% CI
r2
PCE (KB-1) TCE (KB-1) cDCE (KB-1) VC (KB-1) TP-butyric acid TP-ethanol
methanol methanol methanol methanol butyric acid ethanol
-5.5 -13.8 -20.4 -22.4 -5.2 -2.7
( 0.8 ( 0.7 ( 1.2 ( 1.8 ( 0.3 ( 0.9
0.51 0.98 0.94 0.91 0.96 0.67
a Small 95% confidence intervals demonstrate the high level of reproducibility of the data. Correlation factors (r2) are calculated for the Mariotti regression. High r2 values indicate agreement with the Rayleigh model. The low r2 values for KB-1 PCE and TP-ethanol indicate these consortia do not follow Rayleigh type behavior.
Experiment 1 was carried out using subcultures from the same microbial consortium, the same medium formulation, and electron donor as the experiments published by Bloom et al. (9). However, there are significant differences in the results obtained in these two studies. The results of this study for TCE, cDCE, and VC are reproducible (within 95% confidence intervals) over four replicates in contrast with the variation between replicates observed by Bloom et al. (9). In addition, enrichment factors determined in this study are significantly more negative than those reported in Bloom et al. (9). The reasons for the observation of different behavior by the same microbial consortium grown in the same defined VOL. 35, NO. 5, 2001 / ENVIRONMENTAL SCIENCE & TECHNOLOGY
9
905
TABLE 3. Range of Reported Rayleigh Enrichment Factorsa compound PCE TCE cDCE VC
no. of consortia
electron donor
range of reported Rayleigh E
3 3 1 1
MeOH, EtOH, butyric acid MeOH, lactate MeOH MeOH
-5.21 -2.5 (9), -6.6 (9), -7.1 (8), -13.81 -14.1 (9), -16.1 (9), -20.41 -21.5 (9), -22.41, -26.6 (9)
a Overall observed range in is shown as well as individual and their sources (1 refers to this study, other italicized numbers in parentheses refer to reference list). Enrichment factors are not included for experiments which did not follow Rayleigh behavior. Estimates of by Hunkeler et al. (10) are not included as they were not calculated by applying a Rayleigh regression.
mineral medium, with the same electron donor, are not clear. It may have been due to small differences in the experimental set up between the two studies, or other factors. However, the observation of variation during reductive dechlorination by the same consortium under ostensibly the same conditions does demonstrate the need for further study to understand the factors controlling isotopic fractionation during reductive dechlorination. Non-Rayleigh Fractionation during Biodegradation of PCE. For KB-1 (Figure 3a) and TP-ethanol (Figure 3e), where the PCE data falls off the Rayleigh curve in the latter half of the experiments, the 95% confidence intervals on the enrichment factors are still small, indicating that the data are highly reproducible (Table 2). However, the r2 correlation coefficients for these regressions are low indicating that these data cannot be described by the Rayleigh model (Table 2). For all characterizations of carbon isotopic fractionation during reductive dechlorination, only experiments with PCE have shown such non-Rayleigh fractionation behavior. There is no clear explanation for the observation of Rayleigh behavior in one experiment (TP-butyric PCE) and non-Rayleigh behavior in two other experiments (TP-ethanol PCE, KB-1 PCE). The fact that this behavior is observed solely in experiments with PCE suggests that this effect may be due to some factor related to the properties of PCE. It is possible that the low solubility or the high Kow of PCE may result in local substrate limitation and a change in the extent of fractionation observed. Alternatively, this behavior may be due to toxicity effects, nutrient limitation, or variation in reaction pathways over the course of degradation. This observation is important because until the factors causing this difference in behavior for PCE are understood, it cannot be assumed that dechlorination of other chlorinated ethenes will necessarily exhibit Rayleigh behavior at all times. Implications to Field Applications. Notwithstanding the variation observed in enrichment factors reported to date, there are consistent trends in the isotopic fractionation observed during microbial reductive dechlorination of the chlorinated ethenes. The trend of decreasing extent of fractionation with increasing chlorination observed for the KB-1 consortium is consistent in studies with other microbial consortia. In general, for reductive dechlorination, the magnitude of the enrichment factor () decreases with increasing chlorination of the chlorinated ethene (Table 2). This observation, that enrichment factors observed to date fall into distinct ranges, implies that the relative extent of degradation of a chlorinated ethene in the field can potentially be qualitatively estimated using the range of enrichment factors shown in Table 3. Such an approach is well established and widely used in oil and gas research. Despite ranges in enrichment factors of 10 to >20‰ for carbon isotope fractionation and 95 to 285‰ for hydrogen isotope fractionation, a pronounced isotopic enrichment in residual CH4 is used as a definitive indicator of the effects of microbial oxidation (e.g. ref 24). Similarly, the consistent observation of isotopic fractionation during microbial reductive dechlorination of the chlorinated ethenes means that isotopic analysis may be useful as a tool to identify the occurrence 906
9
ENVIRONMENTAL SCIENCE & TECHNOLOGY / VOL. 35, NO. 5, 2001
of biodegradation. Though it is not always able to be described by a Rayleigh model, isotopic fractionation results in significant enrichment of the residual carbon pool over the course of degradation, even to the extent that positive isotopic values are obtained for residual contaminants. Given that the known isotopic signatures for all neat chlorinated ethenes are more negative than -20‰, the observation of much more highly enriched δ13C values is a good indication of the occurrence of biodegradation.
Acknowledgments The authors wish to thank the organizations which contributed funding for this project - The Natural Sciences and Engineering Research Council of Canada and the University Consortium on Solvents-in-Groundwater. Thanks are also due to N. Arner, H. Li, C. Heidorn, V. Kaseros, D. Kostenberg, and K. Krastle for technical expertise and analytical support. Thank you to E. Cox and D. Major from GeoSyntec. Finally, thank you to three anonymous reviewers for their constructive comments.
Literature Cited (1) Madsen, E. L. Environ. Sci. Technol. 1991, 25, 1663-1673. (2) Wiedemeier, T. H.; Swanson, M. A.; Wilson, J. T.; Kampbell, D. H.; Miller, R. N.; Hansen, J. E. Ground Water Monitoring Rev. 1996, Summer, 186-194. (3) Gillham, R. W.; O′Hannesin, S. F. Ground Water 1994, 32, 958967. (4) Maymo-Gatell, X.; Chein, Y.; Gossett, J. M.; Zinder, S. H. Science 1997, 276, 1568-1571. (5) Neumann, A.; Scholz-Muramatsu, H.; Diekert, G. Archives Microbiol. 1994, 162, 295-301. (6) Wild, A.; Hermann, R.; Leisinger, T. Biodegradation 1996, 7, 507-511. (7) Aggarwal, P. K.; Fuller, M. E.; Gurgas, M. M.; Manning, J. F.; Dillon, M. A. Environ. Sci. Technol. 1997, 31, 590-596. (8) Sherwood Lollar, B.; Slater, G.; Ahad, J.; Sleep, B.; Spivack, J.; Brennan, M.; MacKenzie, P. Org. Geochem. 1999, 30, 813-820. (9) Bloom, Y.; Aravena, R.; Hunkeler, D.; Edwards, E.; Frape, S. K. Environ. Sci. Technol. 2000, 34, 2768-2772. (10) Hunkeler, D.; Aravena, R.; Butler, B. J. Environ. Sci. Technol. 1999, 33, 2733-2738. (11) Slater, G. F.; Dempster, H. S.; Sherwood Lollar, B.; Ahad, J. Environ. Sci. Technol. 1999, 33, 190-194. (12) Harrington, R. R.; Poulson, S. R.; Drever, J. I.; Colberg, P. J. S.; Kelly, E. F. Org. Geochem. 1999, 30, 765-775. (13) Poulson, S. R.; Drever, J. I. Environ. Sci. Technol. 1999, 33, 36893694. (14) Galimov, E. M. The Biological Fractionation of Isotopes; Academic Press: Orlando, 1985. (15) Dayan, H.; Abrajano, T.; Sturchio, N. C.; Winsor, L. Org. Geochem. 1999, 30, 755-763. (16) Slater, G. F.; Dempster, H. D.; Sherwood Lollar, B.; Spivack, J.; Brennan, M.; MacKenzie, P. First International Battelle Conference on Remediation of Chlorinated and Recalcitrant Compounds; Monterey, CA, 1998. (17) Edwards, E.; Cox, E. Fourth International Symposium on In Situ and On-Site Bioremediation, New Orleans, LA, 1997; pp 261-265.
(18) Grbic-Galic, D.; Vogel, T. M. Appl. Environ. Microbiol. 1987, 53, 254-260. (19) Gossett, J. M. Environ. Sci. Technol. 1987, 21, 202-208. (20) Bagley, D. M.; Lalonde, M.; Kaseros, V.; Stasiuk, K. E.; Sleep, B. E. Water Res. 2000, 34, 171-178. (21) Mansour, I.; Sleep, B. E.; Liss, S. N. Environ. Sci. Technol. 1998, 32, 3579-3585. (22) Dempster, H. S.; Sherwood Lollar, B.; Feenstra, S. Environ. Sci. Technol. 1997, 31, 3193-3197.
(23) Mariotti, A.; Germon, J. C.; Hubert, P.; Kaiser, P.; Letolle, R.; Tardieux, A.; Tardieux, P. Plant Soil 1981, 62, 413-430. (24) Whiticar, M. J. Chem. Geol. 1999, 161, 291-314.
Received for review August 15, 2000. Revised manuscript received November 29, 2000. Accepted December 4, 2000. ES001583F
VOL. 35, NO. 5, 2001 / ENVIRONMENTAL SCIENCE & TECHNOLOGY
9
907
