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Dec 3, 2004 - Hydrogen Isotopes during Anaerobic. Biodegradation of MTBE: Microcosm and Field Evidence. TOMASZ KUDER,* , †. JOHN T. WILSON, ‡...
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Environ. Sci. Technol. 2005, 39, 213-220

Enrichment of Stable Carbon and Hydrogen Isotopes during Anaerobic Biodegradation of MTBE: Microcosm and Field Evidence T O M A S Z K U D E R , * ,† J O H N T . W I L S O N , ‡ PHIL KAISER,‡ RAVI KOLHATKAR,§ PAUL PHILP,† AND JON ALLEN† University of Oklahoma, Norman, Oklahoma 73019, United States Environmental Protection Agency, Ada, Oklahoma 74820, and Atlantic Richfield Company, La Palma, California 90623

The conventional approach to evaluate biodegradation of organic contaminants in groundwater is to demonstrate an increase in the concentration of transformation products. This approach is problematic for MTBE from gasoline spills because the primary transformation product (TBA) can also be a component of gasoline. Compound-specific stable isotope analysis may provide a useful alternative to conventional practice. Changes in the δ13C and δD of MTBE during biodegradation of MTBE in an anaerobic enrichment culture were compared to the δ13C and δD of MTBE in groundwater at nine gasoline spill-sites. The stable isotopes of hydrogen and carbon were extensively fractionated during anaerobic biodegradation of MTBE. The stable isotope enrichment factor for carbon (C) in the enrichment cultures was -13 (-14.1 to -11.9 at 95% confidence level), and the hydrogen enrichment factor (H) was -16 (-21 to -11 at 95% confidence level). The isotope enrichment factors for carbon and hydrogen during anaerobic biodegradation indicate that the first reaction is enzymatic hydrolysis of the O-Cmethyl bond. The ratio of H to C was consistent between the enrichment culture and the field site that provided the inoculum, and with the other eight sites, suggesting a common degradation pathway. Compound-specific isotope evidence is discussed in terms of its utility for monitoring in situ biodegradation, in particular, for measuring how much MTBE was degraded. For the studied field sites, significant biodegradation of the original mass of MTBE is suggested, in some cases exceeding 90%.

Introduction Compound-specific stable isotope analysis (CSIA) is a recent addition to the environmental geochemistry toolbox, primarily in the field of volatile organic compounds (VOC) remediation. Its practical value is in providing a cost-effective proxy for contaminant degradation by measuring stable isotope ratios on samples collected in situ, offering an * Corresponding author phone: (405)325-3253; fax: (405)325-3140; e-mail: [email protected]. † University of Oklahoma. ‡ United States Environmental Protection Agency. § Atlantic Richfield Company. 10.1021/es040420e CCC: $30.25 Published on Web 12/03/2004

 2005 American Chemical Society

alternative to microcosm studies. Significant changes of stable isotope ratios due to biodegradation have been measured under laboratory and field conditions for chlorinated solvents (1-8), the BTEX compounds (9-19), crude oil alkyl-benzenes (20), naphthalene (19), phenol (21), n-alkanes (22), and fuel oxygenates (23-26). Chemical transformation of some of the same compounds was shown to cause similar changes in isotope ratios (27), but other processes contributing to attenuation (such as dispersion, sorption, volatilization, and phase partitioning produce) exhibited minimal effects on the isotope ratios (28-32). Methyl t-butyl ether (MTBE) is rapidly mineralized by a variety of aerobic bacteria (33). From the practical point of view, the bulk of MTBE plumes are anaerobic, and anaerobic biodegradation will be more significant for natural attenuation. A number of studies have shown MTBE degradation in anaerobic microcosms and in the field, under a range of electron acceptor conditions (34-42). Degradation under anaerobic conditions (43, 44) is slower and in most cases may be restricted to biotransformation of MTBE to t-butyl alcohol (TBA). As a consequence, natural attenuation of anaerobic plumes tends to be relatively slow, resulting in the accumulation of t-butyl alcohol (TBA) as MTBE disappears. While microcosm studies offer easily interpretable results, they are expensive and tend to be time-consuming. In particular, microcosm studies may take from several months to over a year to complete. Such studies can only show that the aquifer harbored microorganisms that were capable of degrading MTBE under the conditions that pertained at the time the aquifer material was sampled. The stable isotope approach provides direct information on the extent to which MTBE has been degraded in the groundwater. The application of monitored natural attenuation to MTBE at gasoline spill sites is limited by the difficulty to convincingly demonstrate the biological removal of MTBE. Trends in the concentration of MTBE over time remain the chief line of evidence for attenuation. Evidence of biodegradation based on concentrations of electron acceptors and degradation products is typically inconclusive. The presence of TBA in groundwater is not conclusive evidence for biodegradation of MTBE because TBA is also a trace component of commercial MTBE added to gasoline. TBA is miscible with water and will partition from gasoline, potentially reaching high concentration in groundwater (45) in the absence of MTBE biodegradation. The objective of this work was to document anaerobic biodegradation of MTBE using CSIA of both carbon and hydrogen. Our initial publication reported enrichment of carbon isotopes during anaerobic biodegradation of MTBE (26). Recent progress in method optimization permits analysis of hydrogen isotope ratios in MTBE in aqueous samples at environmentally relevant concentrations. The immediate goals of the current work were to (i) collect data from enrichment cultures and from a series of anaerobic field sites on the fractionation of carbon and hydrogen isotopes during anaerobic biodegradation of MTBE; (ii) use the fractionation patterns of carbon and hydrogen to evaluate the most likely biochemical mechanisms of anaerobic degradation; and (iii) evaluate the utility of carbon and hydrogen isotope data for management of contaminated sites. Having data on both carbon and hydrogen is particularly powerful. As will be demonstrated, the concurrent fractionation of C and H can prove that a field sample was degraded through the anaerobic pathway. If the pathway was anaerobic, then the isotope data can be used to calculate the extent of biodegradation. The calculated extent of biodegradation can be compared to VOL. 39, NO. 1, 2005 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

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TABLE 1. Summary Information on Field Sites field site site code no. of wells studied date collected MTBE spill date remediation activities historical max MTBE (µg/L; yr) current max MTBE (µg/L; yr) range of MTBE δ13C (‰) range of MTBE δD (‰) Fbio by eq 3b

OC 1

OC 2

OC 3

OC 4

OC 5

6 10/2002 pre-1996 dual phase extraction 550 000 (7/2001) 17 600 (4/2003) -27.9 to -13.3 -78 to -60 NBDc to 0.33

8 10/2002 pre-1985 dual phase extraction 100 000 (4/2001) 235 (8/2003) -29.0 to 9.4 -95 to -38 NBD to 0.05

6 10/2002 pre-1988 dual phase extraction 110 000 (6/1999) 1 400 (8/2003) -27.0 to 24.0 -77 to -9 NBD to 0.012

4 10/2002 pre-1987 dual phase extraction 200 000 (9/2000) 94 000 (10/2003) -31.8 to -12.4 -85 to -66 NBD to 0.31

5 12/2002 pre-1996 soil vacuum extraction 200 000 (5/1997) 1 800 (12/2002) -28.9 to 27.1 -83 to -78a NBD to 0.01

field site site code

OC 6

OC 7

OC 8

PNJ

no. of wells studied by CSIA date collected MTBE spill date remediation activities historical max MTBE (µg/L; yr) current max MTBE (µg/L; yr) range of MTBE δ13C (‰) range of MTBE δD (‰) Fbio by eq 3b

7 12/2002 pre-1996 soil vacuum extraction 4 000 000 (12/1998) 246 000 (12/2002) -30.6 to 56.8 -88 to -49a NBDc to 0.001

3 3/2003 pre-2000 none 81 000 (7/2000) 79 (3/2003) -20.5 to 57.8 -67 to 42 0.6 to 0.001

3 3/2003 pre-1996 none 770 000 (10/1996) 13 000 (3/2003) -13.1 to -0.7 -59 to -44 0.32 to 0.12

6 7/2001, 7/2002 pre-1993 none 2 200 (9/1994) 251 (7/2002) -26.6 to 5.5 -81 to -24 NBD to 0.08

a δD not available for the samples with most positive δ13C. b F calculated for  ) -12.2, δ13C ) -26.5. See text for explanation. c NBD means 0 no evidence of biodegradation compared to the normal range of δ13C of MTBE in gasoline.

overall attenuation in concentrations to evaluate the relative contribution of biodegradation. To the best of our knowledge, this the first report of the analysis of stable hydrogen isotopes in organic compounds in water when the compounds were present at concentrations in the lower tens of µg/L. This low analytical limit makes the technique viable for environmental samples.

Experimental Procedures Basic Concepts. Depending upon the source and release history, contaminant molecules have a particular 13C/12C and D/1H ratio. This ratio is reported using delta (δ13C or δD, given in per mil units, ‰) notation:

δ13C ) (Rsample/Rstandard - 1)1000

(1)

Rsample and Rstandard represent 13C/12C ratios of the sample and the international standard (VPDB or Vienna Pee Dee Belemnite for δ13C), respectively. It is normally observed that physical and chemical reactions produce changes in the isotope composition between the pools of substrate and the pools of product. Chemical reactions proceed faster for molecules containing the lighter isotopes (e.g., 12C), resulting in reaction products enriched in 12C, while the remaining substrate is enriched in 13C. In the case of hydrogen, the situation is identical, the isotopic species being the lighter 1H and the heavier D. The observed changes are referred to as isotopic fractionation. Isotopic fractionation may be described by the Rayleigh distillation equation (eq 2), where δ13Ct is δ13C of MTBE at time t, δ13C0 is δ13C of MTBE at time t ) 0, C is the carbon isotopic enrichment factor, and F represents the ratio of MTBE concentrations at time ) t and time t ) 0. Rt and Ro are isotope ratios (e.g., 13C/12C) at time t and time t ) 0:

1000 ln{(10-3 δ13Ct + 1)/(10-3 δ13C0 + 1)} ) C ln F (2) Equation 2 is equivalent to eq 2a, and the difference is in isotope ratio notation: 214

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1000 ln(Rt/Ro) )  ln F

(2a)

The benefits of using the Rayleigh model in biodegradation studies are 3-fold: (i) it improves interpretation of field data permitting one to distinguish between isotopic changes due to degradation and source fluctuation; (ii) it allows quantitative conclusions or quantitative boundaries on the rates and extent of elapsed degradation; and (iii) it provides an objective parameter () for comparison between sites or microcosm experiments. Study Sites and Sampling. Groundwater samples were collected from nine sites (Table 1), including eight gasoline stations from Orange County, California, and one gasoline service station in New Jersey (cf. Parsippany, NJ, site; description in ref 26). The study sites in Orange County were selected because they had unexpectedly high concentrations of TBA, which might have been produced from biodegradation of MTBE. Historical concentrations of MTBE are summarized in Table 1. The historical MTBE concentration data were provided by site operators (determined by EPA method 8260 or 8260Bspurge-and-trap with GC-MS). Current MTBE and TBA concentration data were acquired at the University of Oklahoma by PT-GC-MS (modified EPA 524.2 method) or were determined by PT-GC-IRMS analysis, based on mass 44 peak intensity response (see the CSIA method section). Quantitative calibration line was considered adequate for (25% standard spike recovery (GC-MS) and (15% recovery (GC-IRMS), respectively. The complete sample list with numerical data (concentrations and isotope ratios) is provided in Table 1 of the Supporting Information. Samples for analysis of stable isotopes were collected in 40 mL VOA vials, preserved with trisodium phosphate (TSP), and refrigerated until analysis. Enrichment Cultures. Enrichment cultures were inoculated with sediment from a microcosm experiment that exhibited anaerobic biodegradation of MTBE (26). Sediment from several microcosms that completely consumed any MTBE that was originally present was blended together and then used to inoculate the enrichment cultures. The primary

enrichment cultures were constructed in glass serum bottles with a capacity of 25 mL. Each microcosm was inoculated with 2.5 g of sediment. The remaining capacity was filled with a minimal medium dosed with 100 mg/L MTBE. A minimal medium had a bicarbonate buffer. Three of the primary enrichment cultures that showed MTBE degradation after 10 months were used to create secondary enrichment cultures. The secondary enrichment cultures were constructed with 5 mL of medium from the parent enrichment culture and 45 mL of media dosed with 100 mg/L MTBE leaving about 12 mL of headspace in the vial. The headspace was filled with 2-7% v/v hydrogen in nitrogen. The enrichment cultures were prepared, sealed, and stored upside down in a glovebox, under an atmosphere containing 2-5% v/v hydrogen and less than 1 ppmv oxygen and incubated at room temperature (20-22 °C). The day before sampling, the cultures were shaken and inverted. The solids were allowed to settle, then the caps were removed, and 1 mL of medium was removed and diluted with 13 mL of distilled water containing 0.14 g of trisodium phosphate as a preservative. The complete microcosm sample list with numerical data (concentrations and isotope ratios) is provided in Table 2 of the Supporting Information. Stable Isotope Analytical Techniques. Stable isotope results were obtained by a combined purge-and-trap (PT) gas chromatograph-isotope ratio mass spectrometer (GCIRMS). This technique permits analysis of low concentrations of aqueous VOC without the need to process high volumes of water (26, 46), allowing analysis of samples collected by routine monitoring water sampling or aliquots of microcosm cultures. Analytes were extracted from water by purge and trap (PT model 4560 by O&I Analytical). PT effluent was analyzed with a GC-IRMS instrument with a thermal conversion interface (Finnigan MAT 252 IRMS for the carbon analyses, Finnigan Delta XL for hydrogen). The δ13C of the samples was determined with a precision of 0.5‰ or better. The δD samples were determined with a standard deviation no greater than 10‰ (median standard deviation 3‰). Details of the analytical protocols are provided in the Supporting Information.

Results and Discussion Stable Isotope Fractionation in Anaerobic Enrichment Cultures. Previously published results (26) for anaerobic MTBE biodegradation in microcosms constructed with material from the Parsippany, NJ, site show that (i) the MTBE concentration decreased in live but not in autoclaved microcosms; (ii) the TBA concentration increased in the microcosms where MTBE disappeared; (iii) carbon isotope fractionation of MTBE occurred only in live microcosms; and (iv) the anaerobic biodegradation process resulted in extensive fractionation of the stable carbon isotopes as compared to aerobic biodegradation. Figure 1 shows the progress of carbon and hydrogen isotope fractionation during anaerobic biodegradation of MTBE in microcosms constructed with sediment from the Parsippany, NJ, service station (26) and in enrichment cultures derived from the microcosms. The magnitude of carbon isotope effects in the enrichment cultures was similar to those published before for the microcosms (Figure 1A) and strongly differed from the values reported for aerobic biodegradation (25). The range of fractionation of hydrogen isotopes in the anaerobic enrichment cultures correlated with the extent of biodegradation (Figure 1B). The extent of hydrogen isotope fractionation during anaerobic biodegradation is lower than the range of the fractionation reported by Gray et al. (25) for aerobic biodegradation. Characterization of both C and H enrichment factors provides a two-dimensional characterization of a chemical reaction, allowing discrimination or correlation of samples

FIGURE 1. Enrichment of stable isotopes of carbon (A) and hydrogen (B) in laboratory experiments. Diamonds are data from microcosms constructed using sediment from a site in Parsippany, NJ (shown after ref 26). Circles are enrichment cultures developed from the microcosms. The value of F is the fraction of MTBE remaining after biodegradation. Vertical error bars represent error in determination of the isotope ratio (δ13C error bar does not exceed the diameter a data point). Horizontal error bars represent error in determination of the concentration of MTBE. Solid lines are shown for reference and represent the reported values for aerobic bacteria (25). based on the mechanism of degradation. A simplified approach based on regression of δ13C and δD (XY plot where the axes represent measured delta values of carbon and hydrogen) does not provide  values; instead, the regression slope represents the ratio of H/C (derived from simplified Rayleigh equation of Mariotti et al.; 47). The insert in Figure 2 compares the fractionation of hydrogen isotopes (δD) to the fractionation of carbon isotopes (δ13C) in the enrichment cultures. There was a strong linear correlation between δ13C and δD (R2 ) 0.86). The solid lines in Figure 2 depict the expected values of δ13C and δD following the pattern for aerobic biodegradation published by Gray et al. (25). The H/C is very different from the relationship that would be expected from aerobic biodegradation. Enrichment factors () were calculated relative to the MTBE concentration (eq 2), using a collection of replicate enrichment cultures sampled as biodegradation progressed. In the present study, the value obtained for C was -13 (-14.1 to -11.9 at 95% confidence). The estimate of C is not statistically different from the previously reported estimate of -8.2 (-11.3 to -5.2 at 95% confidence level; recalculated by eq 2 from data in ref 26); however, the confidence interval on estimate of C in the present study is better than in the previous study ((1.1 as compared to (3.1). The regressions VOL. 39, NO. 1, 2005 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

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FIGURE 2. Concurrent fractionation of hydrogen isotopes and carbon isotopes during anaerobic biodegradation of MTBE. The insert shows microcosm data. were calculated for fixed intercept (known from direct measurement of isotope ratios of the undegraded substrate). In comparison, the fractionation of carbon during aerobic MTBE biodegradation is less extensive. Hunkeler et al. (23) and Gray et al. (25) reported values for the C in the range between -1.4 and -2.4 during biodegradation of MTBE by aerobic bacteria. In the present study, the value obtained for H during anaerobic biodegradation was -16 (-21 to -11 at 95% confidence). Gray et al. (25) reported values of H that varied from -30 to -69. Stable Isotope Fractionation in the Field. MTBE collected from monitoring wells at nine gas stations show significant isotopic enrichment (Table 1) with values as positive as +58‰ δ13C and +42‰ δD being observed. The isotopic enrichment factors for MTBE biodegradation in the microcosms can be extrapolated to the field samples if it can be shown that the same physiological process was responsible for MTBE biodegradation in the microcosms and in the field samples. Figure 2 compares δD against δ13C in the field samples. As was the case with the enrichment cultures, the field samples show a linear correlation between δ13C and δD (R2 ) 0.92). The slope of the regression of δD on δ13C for the field data was 1.3 (1.2-1.4 at 95% confidence), while the slope for the enrichment cultures was 1.6 (1.2-1.9 at 95% confidence). The concurrent fractionation of hydrogen and carbon isotopes in the field samples cannot be distinguished from the concurrent fractionation in the anaerobic enrichment

culture, but the concurrent fractionation during anaerobic biodegradation differs strongly from all published data for aerobic MTBE biodegradation. On the basis of this evidence, anaerobic biodegradation is the dominant process in each of the field samples studied, and the contribution from aerobic degradation is at best marginal. MTBE Stable Isotope Compositions at Gas Station Sites. The range of δ13C in commercial MTBE measured to date is between -27.4 and -33‰, based on two surveys of commercial gasolines (12 samples from the U.S., published by Smallwood et al. (48); samples from 24 countries in Europe, Southeast Asia, and South America, published by O’Sullivan et al. (49)). The δD of commercial MTBE is less well-defined. To date, δD values in several commercial MTBE samples and from the source zones of MTBE plumes vary from -80 to -125‰. The values for δ13C in MTBE in the groundwater (Table 1) are clearly more positive than values expected for commercial MTBE and can only be explained by biodegradation of the MTBE in groundwater. While other processes such as sorption, equilibrium phase partitioning, or volatilization may fractionate isotopes, the magnitude of those changes is small (28-32). The extent of fractionation is generally from 1 to 2 orders of magnitude less than fractionation expected during MTBE biodegradation. When the value of δ13C in MTBE in the field samples is compared to the concentration of MTBE remaining in the groundwater, the samples with the lowest concentrations of MTBE tend to have the most positive value for δ13C (Figure 3). These are presumably the samples with the most extensive degradation. In many of the wells, the distribution of δ13C against the concentration of MTBE roughly approximates the Rayleigh fractionation model for biodegradation. However, in many other wells, the value of δ13C is much lower, indicating that dilution and dispersion made major contributions to the reduction in concentration of MTBE in these wells. At several sites (OC 1, OC 2, OC 3, OC 4, OC 5, OC 6, and PNJ in Table 1), at least one of the wells showed little or no evidence of fractionation of 13C in MTBE. The δ13C for MTBE was in the range expected for MTBE in gasoline or only slightly heavier. In other wells at the same sites the δ13C for MTBE was up to 84‰ heavier than the normal range of gasoline. The extent of biodegradation of MTBE, as inferred from δ13C for MTBE, was very heterogeneous at the field sites. Reaction Mechanisms. Stable isotope analysis has been used extensively to characterize enzymatic reaction mechanisms (50). The isotope fractionation during anaerobic degradation of MTBE can be used to infer the mechanism of the initial biochemical reaction. Finneran and Lovley observed that TBA is the first major product of anaerobic biodegradation of MTBE (40). Likewise, significant amounts

FIGURE 3. Relationship between MTBE concentration and δ13C at the nine field sites. Arrows indicate expected trends in isotope composition change due to biodegradation as opposed to trends due to dilution and dispersion. 216

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of TBA accumulated during anaerobic degradation of MTBE in the enrichment cultures (Table 2 in the Supporting Information). The necessary first step in the production of TBA from MTBE is cleavage of the bond between the ether oxygen and either the t-butyl group or the methyl group in MTBE. On the basis of data published by Huskey (51), the value of C that would be expected from the cleavage of a single C-O bond in a molecule with five carbon atoms is -12.2. The value of C calculated for the enrichment cultures (-12.9 ( 1.1 at 95% confidence) is in good agreement with the value of  that would be expected for cleavage of a C-O bond. This value of C is higher than isotope effects that would be expected from cleavage of C-H bonds. The potential for significant primary isotope effects is limited to the atoms of the methyl group and to the central carbon atom of the t-butyl group. A measurable carbon isotope effect on the reaction product TBA would be expected if the actual reaction involved cleavage of the bond between oxygen and the t-butyl group. There was no evidence in the anaerobic enrichment cultures of a shift in the δ13C of TBA during biodegradation of MTBE (Table 2 in the Supporting Information). Cleavage of the bond between oxygen and the t-butyl group cannot make a significant contribution to biodegradation of MTBE in the enrichment cultures. The only alternative is cleavage of the bond between oxygen and the methyl group. The fractionation of hydrogen isotopes during anaerobic MTBE biodegradation supports the conclusions based on fractionation of carbon. Because only a few enrichment cultures were available, the value of H is estimated with low precision (-21 to -11 at 95% confidence level). However, if the slope of δD/δ13C (equivalent to the ratio of C/H) in the larger set of the field data is considered, a narrower range of H is suggested by extrapolating from C (-20 to -14 at 95% confidence level). Even so, the value of H is roughly equivalent to C. Secondary kinetic effects accompanying C-O cleavage and/or equilibrium effect(s) preceding the cleavage can be proposed to account for the δD fractionation. If the reaction proceeded through cleavage of a hydrogen bond, the expected value of H would be much higher. The cleavage of a C-O bond is in agreement with degradation of MTBE by hydrolase enzymes similar to chymotrypsin or lysozyme as discussed by O’Reilly et al. (52). On the other hand, the absence of strong hydrogen effect is not consistent with biodegradation through any reaction where the rate-limiting intermediate might be formed by cleavage or substitution of a C-H bond. Using δ13C to Predict the Fraction of MTBE Remaining after Biodegradation at Field Scale. As indicated by the loglinear correlation of the fraction of MTBE remaining after biodegradation to the stable carbon and hydrogen isotope ratios (Figure 1), anaerobic biodegradation of MTBE in microcosms experiments is a Rayleigh-type process and can be mathematically described by the Rayleigh equation (eq 2). As discussed previously, this equation permits calculation of the amount of MTBE lost to biodegradation, as opposed to the overall attenuation due to the combined contribution of biodegradation (if any), dilution, sorption, volatilization, etc. The extent of biodegradation will be calculated from the carbon isotope data instead of hydrogen isotope data because there is better analytical precision of  for carbon isotopes (C) and a narrower range of δ13C0 in gasoline. The fraction of MTBE remaining after biodegradation as predicted by the isotope analysis (Fbio) is obtained from eq 3:

MTBE isotope composition measured for a given field sample is δ13Ct. The selection of the remaining two parameters, C and δ13C0, has to be discussed in greater detail. To ensure that the calculated value of Fbio does not overestimate the extent of MTBE biodegradation, which would undermine the significance of the stable isotope approach for environmental remediation studies, it is necessary to either provide accurate values for the two parameters or to use conservative estimates based on indirect references.

Fbio )

To provide a second benchmark, the fraction remaining after biodegradation was calculated following eq 5, where CMTBE and CTBA are the measured concentration of MTBE or TBA

exp(1000 ln{(10

-3

13

δ Ct + 1)/(10

-3

13

δ C0 + 1)}/C) (3)

Although all of the experimental values for the anaerobic enrichment factor () that have been measured to date have been measured on microcosms developed from a single field site, the experimental values of C are statistically indistinguishable from the value that would be theoretically expected for the maximum isotope effect for C-O bond cleavage (see previous section). We propose that calculation of Fbio should be based on the theoretical value C ) -12.2, unless sitespecific microcosm data will indicate otherwise. Since C ) -12.2 represents the largest feasible isotope effect, calculations based on this value may possibly underestimate but will not overestimate the extent of biodegradation. The initial value of the isotopic composition of contaminant can be based on a direct measure of the isotopic composition of source material collected in situ. This approach has been successful in a number of studies (5, 16-19). On the other hand, at many MTBE sites it is no longer possible to collect a sample of the gasoline that was originally spilled. A single plume of MTBE may originate from several different spills of gasoline, which may have occurred at different times. If the gasoline that was spilled is different, it is possible that the initial isotopic composition of the MTBE in the gasoline was different. Unless it can be proven that the isotopic composition of MTBE in the gasoline that was spilled is homogeneous, it is possible that MTBE in residual gasoline in different parts of the source area has a different isotopic composition. It would be safe to assume that a single release at one point in time was homogeneous but a slow release over a period of years may not be homogeneous. To deal with this uncertainty, it is better to use the most positive value ever reported for δ13C of MTBE in gasoline as δ13C0 in eq 3 rather than some low value of δ13C measured in MTBE in groundwater at the site. On the basis of the two available commercial gasoline surveys (48, 49), and providing a safety margin for analytical precision, it is proposed that a value of δ13C0 ) -26.5 be used in eq 3 in situations where the direct measurement of δ13C0 is not possible. To provide a benchmark for the estimate of biodegradation based on fractionation of 13C in MTBE, the extent of attenuation of MTBE at field scale was estimated from conventional monitoring data at the site. Following eq 4, the attenuation of MTBE (fconc refers to the currently observed fraction of initial MTBE concentration) was calculated from the measured concentration of MTBE in a particular well at the time of sampling (CMTBE) and the historical maximum concentration of MTBE at the site (CMTBEmax). Equation 4 assumes that CMTBEmax was the initial concentration before biodegradation. Using the maximum value for a given site to evaluate all data from that site, rather than using the local maximum from each specific monitoring well, is intended to account for any attenuation of MTBE concentration between the source and the monitoring well:

fconc ) CMTBE/CMTBE max

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(4)

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FIGURE 4. Fraction of MTBE remaining after biodegradation (Fbio) as predicted from the δ13C in MTBE using eq 3 as compared to the attenuation of MTBE concentration by eq 4 and the fraction remaining as predicted from concentrations of MTBE and TBA by eq 5. Diamonds are field data predicted from eq 4. Open circles are field data predicted from eq 5. Triangles are microcosm data predicted from eq 4. at the time of sampling, and MWMTBE and MWTBA are the molecular weights of MTBE and TBA (88 and 74, respectively):

Fbio ) CMTBE/(CMTBE + CTBAMWMTBE/MWTBA)

(5)

Equation 5 calculates the concentration of MTBE that would be necessary to produce the amount of TBA measured in the groundwater. It assumes that all of the TBA in water came from biodegradation of MTBE. Equation 5 further assumes that none of the TBA in groundwater was biodegraded and that the effects of dilution, dispersion, and sorption change the concentrations of TBA and MTBE in the same proportion. Figure 4 compares the extent of MTBE biodegradation in the field samples as estimated from CSIA data to the extent of attenuation calculated by eqs 4 and 5. The microcosm results are shown for comparison. As would be expected, the extent of biodegradation in the microcosm studies as estimated from CSIA is in good agreement with the extent estimated by removal of MTBE and production of TBA. Figure 4 shows that the estimate of the extent of biodegradation based on stable isotope fractionation of MTBE was conservative as compared to the two estimates based on the concentrations of MTBE and TBA. The estimate of biodegradation based on CSIA was never significantly greater than the two estimates based on concentrations of MTBE and/or TBA. MTBE attenuation predicted by eq 4 does not discriminate between concentration decrease due to biodegradation, dilution in a well, dispersion along a flow path, or sorption; therefore, the reduction of MTBE concentration tends to exceed the decrease of MTBE mass. There are several reasons why the extent of biodegradation predicted by eq 5 would be greater than the true extent of biodegradation. At least some TBA was probably present in the gasoline that was spilled and partitioned to the groundwater with the MTBE. More MTBE than TBA would be lost from groundwater due to volatilization, and MTBE would partition to aquifer matrix material more readily than TBA. There are also reasons why the extent of biodegradation predicted by eq 3 from CSIA could be less than the true extent. The approach used herein, where conservative value for the source isotope composition is applied, almost certainly results in the calculated Fbio values being underestimated. 218

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However, even for the most negative reference value of δ13C0 ) -33‰, the Fbio obtained by eq 3 remains more conservative than the results from eqs 4 and 5 (not shown). If the groundwater was in contact with residual gasoline, the fractionated MTBE might be diluted with fresh MTBE that partitions from the gasoline. If biodegradation in a plume is heterogeneous, where MTBE in one portion is more extensively degraded than another portion, the MTBE in water from a monitoring well that sampled both portions would be dominated by MTBE from the portion that did not degrade as extensively. It is also theoretically possible that eq 3 would underestimate the extent of biodegradation because biodegradation proceeded through the aerobic pathway (aerobic MTBE degradation results with much less isotope fractionation). Consideration of combined carbon and hydrogen data (cf. Figure 2) eliminates this possibility for the nine sites presented in the current paper. Compound-Specific Isotope Analysis for Evaluation of Contaminated Sites. The literature on aerobic MTBE degradation (25) and aerobic and anaerobic biodegradation of BTEX (12-15, 18) suggests that hydrogen fractionation provides more robust evidence of biodegradation. The limited fractionation of carbon often rendered δ13C inconclusive. This is largely true because these reactions proceed through a rate-limiting step that involves cleavage of a C-H bond. In the case of anaerobic biodegradation of MTBE, the numerical difference between carbon and hydrogen fractionation magnitude is not as high. The utility of hydrogen data is further limited by lower analytical precision for δD and by a wider range of values of δD in MTBE in gasoline. Although hydrogen isotope data are invaluable whenever it is necessary to distinguish between anaerobic and aerobic biodegradation, analysis of carbon stable isotopes appears to be the better choice for the purpose of both qualitative and quantitative study of anaerobic MTBE biodegradation. Quantitative compound analysis of carbon stable isotopes can provide an estimate of the fraction of MTBE that was biologically degraded, as opposed to the combined contribution of dilution, dispersion, or sorption and biodegradation. The utility of this approach has been tested at several field sites with BTEX and naphthalene contamination (17, 19), where changes in contaminant concentration could be

quantitatively explained by biodegradation based on the Rayleigh isotope fractionation model. Several previous studies have followed changes in stable isotopes in monitoring wells along an inferred flow path in groundwater. For this approach to be creditable, there must be good coverage of the site with monitoring wells, and the site must be well-characterized to prove that the wells involved in the comparison do, in fact, lie along a flow path. The conservative approach to estimate anaerobic biodegradation of MTBE presented in this manuscript permits interpretation of data even when site hydrology is not wellunderstood and representative source MTBE cannot be obtained. This is made possible by the large carbon isotope effects during anaerobic biodegradation of MTBE, as compared to biodegradation of other contaminants (90% biodegradation shifts δ13C by 28‰ and 99% biodegradation shifts δ13C by 56‰). It is also made possible by the narrow range of isotopic composition of commercial MTBE (δ13C0 range is only 5‰). At a site where removal of MTBE is dominated by aerobic biodegradation, the extent of biodegradation would be dramatically underestimated. However, a site dominated by aerobic biodegradation can readily be distinguished from a site dominated by anaerobic biodegradation by comparing the fractionation of hydrogen to the fractionation of carbon. While the conservative approach is likely to underestimate the extent of MTBE biodegradation at field scale, it provides a robust conservative boundary on the extent of biodegradation. On the basis of eq 3, a δ13C for MTBE of -18.5‰ or higher would indicate that at least 50% of the original MTBE was biodegraded. A value of δ13C exceeding 0 would indicate that approximately 90% of the MTBE had been degraded. The technique can readily and unequivocally document biodegradation of MTBE. Study sites OC 1 and OC 4 showed the minimum fractionation among the nine sites (Table 1). At these two sites, the maximum value of δ13C for MTBE corresponded to approximately 70% removal (Table 1). At three sites, the maximum removal of MTBE was near 90%, and at four sites the maximum removal was near 99%. A conservative estimate of the extent of natural biodegradation of MTBE in a plume would be useful to an evaluation of monitored natural attenuation or for a risk management of a plume.

Acknowledgments The U.S. Environmental Protection Agency through its Office of Research and Development funded and managed part of the research described here through in-house Task 5857 (Monitored Natural Attenuation of MTBE). It has not been subjected to Agency review and therefore does not necessarily reflect the views of the Agency, and no official endorsement should be inferred.

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Supporting Information Available Details of the compound-specific stable isotope methodology and numerical data for MTBE and TBA concentrations, MTBE and TBA isotope ratios of microcosms, and MTBE isotope ratios of field. This material is available free of charge via the Internet at http://pubs.acs.org.

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Received for review April 29, 2004. Revised manuscript received October 14, 2004. Accepted October 19, 2004. ES040420E