Environ. Sci. Technol. 2009, 43, 2793–2799
Isotopic Fractionation of Methyl tert-Butyl Ether Suggests Different Initial Reaction Mechanisms during Aerobic Biodegradation JENNIFER R. MCKELVIE,† MICHAEL R. HYMAN,‡ MARTIN ELSNER,§ CHRISTY SMITH,‡ DENISE M. ASLETT,‡ GEORGES LACRAMPE-COULOUME,† 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, University of Toronto, Toronto, Ontario, Canada M5S 3B1, Department of Microbiology, North Carolina State University, Raleigh, North Carolina 27695, and Institute of Groundwater Ecology, Helmholtz Zentrum Mu ¨ nchensNational Research Center for Environmental Health, Neuherberg D-85764, Germany
Received November 21, 2008. Revised manuscript received February 9, 2009. Accepted February 11, 2009.
Carbon isotopic enrichment factors (εC) measured during cometabolic biodegradation of methyl tert-butyl ether (MTBE), ethyl tert-butyl ether (ETBE), and tert-amyl methyl ether (TAME) by Pseudonocardia tetrahydrofuranoxydans strain K1 were -2.3 ( 0.2‰, -1.7 ( 0.2‰, and -1.7 ( 0.3‰, respectively. The measured carbon apparent kinetic isotope effect was 1.01 for all compounds, consistent with the expected kinetic isotope effects for both oxidation of the methoxy (or ethoxy) group and enzymatic SN1 biodegradation mechanisms. Significantly, δ13C measurements of the tert-butyl alcohol and tert-amyl alcohol products indicated that the tert-butyl and tert-amyl groups do not participate in the reaction and confirmed that ether biodegradation by strain K1 involves oxidation of the methoxy (or ethoxy) group. Measured hydrogen isotopic enrichment factors (εH) were -100 ( 10‰, -73 ( 7‰, and -72 ( 2‰ for MTBE, ETBE, and TAME respectively. Previous results reported for aerobic biodegradation of MTBE by Methylibium petroleiphilum PM1 and Methylibium R8 showed smaller εH values (-35‰ and -42‰, respectively). Plots of ∆2H/ ∆13C show different slopes for strain K1 compared with strains PM1 and R8, suggesting that different mechanisms are utilized by K1 and PM1/R8 during aerobic MTBE biodegradation.
Introduction Ether-containing compounds such as methyl tert-butyl ether (MTBE), ethyl tert-butyl ether (ETBE), and tert-amyl methyl ether (TAME) were added to gasoline to improve combustion efficiency (1). Early studies suggested that MTBE was difficult to biodegrade due to the presence of a tertiary carbon atom and an ether linkage (2, 3). However, in recent years, biodegradation of MTBE has been documented in laboratory * Corresponding author address: Stable Isotope Laboratory, Department of Geology, 22 Russell St., Toronto, ON, Canada M5S 3B1; phone: (416) 978-0770; fax: (416) 978-3938. † University of Toronto. ‡ North Carolina State University. § Helmholtz Zentrum Mu ¨ nchensNational Research Center for Environmental Health. 10.1021/es803307y CCC: $40.75
Published on Web 03/13/2009
2009 American Chemical Society
microcosm studies under both aerobic and anaerobic conditions (4). In addition, biodegradation of MTBE has been documented at contaminated field sites using compoundspecific isotope analysis (CSIA) (5-9). CSIA allows for the rapid determination of carbon and of hydrogen isotopic signatures of the organic compound of interest. The ratio of the heavy to light isotopes (R ) 13C/12C or 2H/1H) is expressed relative to an international standard: δ (‰))(Rsample/Rstandard - 1) × 1000
(1)
where carbon isotopic values (δ13C) are measured relative to V-PDB and hydrogen isotopic values (δ2H) relative to V-SMOW (10). During biodegradation, bonds containing the light isotopes (e.g., 12C-1H) will react at a slightly faster rate than those containing heavier isotopes (e.g., 13C-1H or 12 C-2H), resulting in isotopic enrichment of 13C or 2H of the remaining compound (10). This isotopic fractionation associated with a bond breakage is known as a primary isotope effect. However, it is important to note that isotopic fractionation can also theoretically occur in atoms located close to the reacting bond, a phenomenon known as a secondary isotope effect (11). While secondary isotope effects are small for carbon and oxygen, hydrogen secondary isotope effects can be significant (12). Isotopic fractionation exhibited during aerobic MTBE (13-15), aerobic ETBE (15), anaerobic MTBE (16, 17), and anaerobic TAME (16) biodegradation has been shown to follow the Rayleigh isotopic enrichment model: Rf/R0)f (ε/1000)
(2)
where Rf is the measured isotopic ratio of a compound at a given fraction remaining (f), R0 is the initial isotopic composition of the compound, and ε is the isotopic enrichment factor (18). Since laboratory-derived isotopic enrichment factors can be used to quantify in situ biodegradation at contaminated field sites (5-8), assessment of the variability of ε under different redox conditions, but also between microbial species, is important. During aerobic microcosm experiments with sediment from Canadian Forces Base Borden and two sites in California, Vandenberg Air Force Base (VAFB) and Port Hueneme, carbon enrichment factors (εC) showed a small range of -1.4‰ to -1.8‰ (9, 13, 14) but hydrogen isotopic enrichment factors were more variable (εΗ of -29‰ and -66‰ for two replicates) (13) (Table 1). It was suggested that the variability in hydrogen isotopic fractionation might be the result of microenvironments that differed with respect to microbial community composition (13). Studies examining aerobic biodegradation of MTBE by pure cultures fall into two general groupings based on their εC and εΗ values. Methylibium petroleiphilum PM1 and Methylibium R8 showed similar carbon (εC of -2.2‰ and -2.4‰, respectively) and hydrogen (εH -35‰ and -42‰, respectively) isotopic fractionation (13, 15) (Table 1). In contrast, Rhodococcus ruber IFP2001 and β-Proteobacterium L108 had εC values less than -0.5‰ and no measurable hydrogen isotopic fractionation (15) (Table 1). Isotopic enrichment factors can vary as a function of the biodegradation mechanism, since they depend on the first irreversible transformation step, typically the first bond breakage (7, 19, 20). The isotopic signatures measured by CSIA reflect the average isotopic composition of the entire reactant molecule, however, and not just the atoms involved in the reaction. To better discriminate between distinct VOL. 43, NO. 8, 2009 / ENVIRONMENTAL SCIENCE & TECHNOLOGY
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TABLE 1. Comparison of Carbon and Hydrogen Isotopic Enrichment Factors (ε) and Apparent Kinetic Isotope Effects (AKIEs), Measured during Aerobic Biodegradation, Abiotic Hydrolysis, and Abiotic Oxidation of MTBE, ETBE, and TAMEa compd and replicate
experiment
ε
C (‰) ( 95% CI
εH (‰) ( 95% CI
AKIEC
AKIE assuming oxidation
AKIE assuming SN1
∆2H/∆13C ( 95% CI
ref
oxidation by permanganate
MTBE
NA
NA
-109 ( 9
>15b
1.2
NA
acid hydrolysis
MTBE
-4.9 ( 0.6
1.03
-55 ( 7
2.9
1.1
11 ( 1
Pseudonocardia K1
MTBE
-2.3 ( 0.2 (n ) 2)
1.01
-100 ( 10 (n ) 2)
14.9
1.2
48 ( 5
ETBE TAME
-1.7 ( 0.2 (n ) 2) -1.7 ( 0.3 (n ) 1)
1.01 1.01
-73 ( 7 (n ) 1) -72 ( 2 (n ) 1)
4.2 4.5
1.2 1.1
49 ( 4 45 ( 4
M. petroleiphilum PM1
MTBE
-2.2 ( 0.1
1.01
-35 ( 2
1.7
1.1
18 ( 3
13
Methylibium R8
MTBE
-2.4 ( 0.1
1.01
-42 ( 4
2.0
1.1
15
15
R. ruber IFP2001
MTBE ETBE
-0.3 ( 0.1 -0.8 ( 0.1
1.00 1.00
no enrichment -11 ( 4
NA 1.2
NA 1.0
NA 10
15
β-Proteobacterium L108
MTBE
-0.5 ( 0.1
1.00
no enrichment
NA
NA
NA
15
ETBE
-0.7
1.00
-13
1.2
1.0
13
MTBE 1 MTBE 2 MTBE 3
-1.5 ( 0.1 -1.4 ( 0.1 -1.8 ( 0.1
1.01 1.01 1.01
NA -66 ( 3 -29 ( 4
NA 2.5 1.7
NA 1.1 1.0
NA 45 ( 2 12 ( 2
13
Borden enrichment culture
MTBE
-1.8
1.01
NA
NA
NA
NA
14
Port Hueneme microcosm
MTBE
-1.4
1.01
NA
NA
NA
NA
9
VAFB microcosms
c
21
this study
a n ) number of experimental bottles, NA ) not applicable, and CI ) confidence interval. b The determined position-specific enrichment factor εreactive position was >-333‰, resulting from the combined effect of a very large primary KIE and pronounced secondary KIE (21). c Each replicate bottle reported individually.
biodegradation mechanisms, apparent kinetic isotope effects (AKIEs) are used (20). The values of AKIE reflect the difference in reaction rates between the light and heavy isotopes in a reaction (e.g., 12k/13k for carbon) and correct the measured isotopic fractionation for dilution by nonfractionating isotopes present at nonreacting positions and intramolecular competition for chemically equivalent reaction sites (20). In addition, hydrogen AKIE values can also be calculated for reactions where nearby hydrogen atoms do not participate directly in the initial reaction (i.e., SN1 and SN2 mechanisms) but undergo fractionation due to secondary isotope effects (20, 21). To distinguish between different MTBE biodegradation mechanisms (i.e., oxidation, enzymatic SN1 or SN2 reactions; Figure 1), Zwank et al. (7) calculated AKIE values for several previously reported εC and εH values for MTBE biodegradation (6, 13). The carbon AKIE for aerobic biodegradation of MTBE by strain PM1 was consistent with expected intrinsic KIEs for both oxidation and SN1 hydrolysis (7) (Figure 1). However, hydrogen isotopic fractionation during MTBE degradation by strain PM1 was smaller than expected for an oxidative mechanism. Rate limitations occurring before the first bond breakage can potentially reduce observed isotope fractionation (22). These rate limitations can include uptake of the substrate by the cell (23, 24) or binding of the substrate to the enzyme (25). Zwank et al. (7) suggested that the AKIE was suppressed by a high affinity of MTBE for the enzyme and quantitative conversion of the substrate with a reduced flux back to undegraded substrate. This masking effect complicates the use of kinetic isotope effects to determine biodegradation mechanisms and may account for the negligible hydrogen and very small carbon isotopic fractionation during MTBE biodegradation by strains IFP2001 and L108 (15). Elsner et al. (21) evaluated 2794
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hydrogen isotopic fractionation during direct chemical oxidation of MTBE by permanganate and measured an εH of -109‰ (Table 1). The isotopic enrichment exhibited during abiotic oxidation is hypothesized to represent an upper limit (intrinsic kinetic isotope effect (KIE)) for hydrogen isotopic fractionation during aerobic biodegradation via the oxidation mechanism and suggests that the values observable during aerobic biodegradation of MTBE (AKIE) could be considerably larger than have been measured to date (21). Recent studies have demonstrated that plots of δ2H versus δ13C of MTBE have distinct slopes for anaerobic biodegradation compared to aerobic biodegradation (6, 7). Such plots can compensate for masking effects associated with ratelimiting steps, since such effects impact both carbon and hydrogen to the same degree, such that the ratio, or slope, does not change (20, 22, 25, 26). The slopes of δ2H versus δ13C for strains PM1 and R8 are very similar, suggesting a similar biodegradation mechanism for these two bacteria (15) (Table 1). Elsner et al. (21) found that the slope δ2H/δ13C for strain PM1 was distinct from that of abiotic SN1 hydrolysis, suggesting that an SN1 mechanism does not occur for this strain (Table 1). The smaller hydrogen isotopic fractionation for strains PM1 and R8 compared to the theoretical KIE ranges for oxidation warrants further investigation to determine whether the smaller εH values are from masking due to steps preceding the breakage of the C-H bond (i.e., transport or enzymatic efficiency) or whether they are due to an alternate biodegradation mechanism. A comparison of δ2H versus δ13C for strains PM1 and R8 to that of a strain with strong evidence (both microbiological and isotopic) of the oxidation mechanism is one way to further investigate this issue. If steps preceding the initial bond breakage cause the suppressed hydrogen isotopic fractionation in strains PM1 and R8, then
FIGURE 1. Possible biodegradation mechanisms of MTBE to tert-butyl alcohol (TBA), expected KIEs reported in the literature, and correction factors for the number of atoms (n), the number of atoms in reactive positions (x), and the number of positions that are in intramolecular competition (z) used to calculate measured AKIEs (7, 20, 21). Dashed lines depict the initial bond breakage. their slopes of δ2H versus δ13C should match those of strains known to utilize the oxidation mechanism. Pseudonocardia tetrahydrofuranoxydans K1 grows on tetrahydrofuran (THF) and cometabolically oxidizes MTBE and ETBE to tert-butyl alcohol (TBA) and TAME to tert-amyl alcohol (TAA). However, THF-grown cells do not further oxidize TBA and only slowly oxidize TAA in the absence of TAME (Smith and Hyman, unpublished results). Similar to another recently described THF-utilizing, MTBE-oxidizing Pseudonocardia strain (27), the MTBE-oxidizing activity in strain K1 is attributed to the THF-dependent expression of THF-monooxygenase (27, 28). In vivo, this enzyme initiates THF oxidation by generating 2-hydroxytetrahydrofuran from a C-H adjacent to an ether bond (28). An analogous reaction with MTBE would involve an initial oxidation of a C-H bond in the methoxy group (Figure 1). ETBE and TAME biodegradation would occur via a similar mechanism involving oxidation of a C-H bond in the ethoxy and methoxy moieties, respectively (29). The objective of this study was to use three isotopic approaches to investigate the isotopic fractionation patterns associated with oxidation of MTBE, ETBE, and TAME by Pseudonocardia strain K1. Specifically, these approaches were (i) comparison of measured carbon and hydrogen AKIEs during biodegradation by strain K1 to the expected instrinsic KIEs for the different reaction mechanisms (Figure 1), (ii) measurement of δ13C for the TBA and TAA daughter products to further substantiate the biodegradation mechanism with strain K1, and (iii) comparison of the slopes of plots of δ2H versus δ13C for K1 to those of Methylibium strain PM1 and the VAFB microcosms measured in a previous study (15).
Experimental Section Biodegradation of Fuel Oxygenates. Biodegradation of MTBE, ETBE, and TAME was conducted in glass vials (600 mL) capped with screw caps fitted with Teflon-coated butyl rubber inserts. A 5 mL aliquot of concentrated cell suspension of Pseudonocardia strain K1 (DSMZ 44239) (30) was added to 95 mL of sodium phosphate buffer for a final volume of 100 mL and a total cell protein content of ∼0.4 mg. Neat MTBE, ETBE, or TAME was added to reaction vials for initial dissolved concentrations of 1000, 750, and 575 mg L-1,
respectively. Samples of the aqueous phase (2 µL) were removed periodically during the biodegradation experiments for analysis of ethers and corresponding tertiary alcohols by gas chromatography. Additional details on cell growth, harvesting, and analysis are provided in the Supporting Information. At selected times, aqueous samples were taken from the reaction vials and injected directly into amber 40 mL VOA vials sealed with Teflon-lined screw caps (I-Chem, New Castle, DE) containing distilled water and trisodium phosphate (1%, w/v) to halt microbial activity. The volume of sample removed from the reaction vial and the amount of distilled water in the VOA vial were adjusted to achieve final concentrations of MTBE, ETBE, and TAME of 10, 7.5, and 5.75 mg/L, respectively. Carbon and hydrogen isotopic values were measured using purge and trap and CSIA (6, 8) as described in the Supporting Information. Total analytical uncertainty, incorporating both accuracy and reproducibility, is (0.5‰ and (5‰ for δ13C and δ2H, respectively (13, 31). Acid Hydrolysis of Fuel Oxygenates. Acid hydrolysis of MTBE, ETBE, and TAME was conducted in glass serum vials (15 mL) crimp sealed with Teflon-coated butyl rubber stoppers. Samples of each neat ether (7-11 µL) were added to buffer (1 mL, 50 mM NaH2PO4 · H2O adjusted to pH 1.0 with 12 N HCl) and incubated in a shaking water bath (150 rpm) at 40 °C. Ether hydrolysis was monitored over time using gas chromatography as described in the Supporting Information. Once the hydrolysis was complete (g98.4% reacted), samples were added to 40 mL VOA vials containing trisodium phosphate (1%, w/v) for a total volume of 40 mL and a 10 mg/L final concentration of TBA or TAA. The initial δ13C of MTBE and the δ13C of TBA at the end of the reaction were measured using purge and trap and CSIA as described in the Supporting Information. Quantification of Enrichment Factors and Apparent Kinetic Isotope Effects. Fractionation factors were calculated by plotting ln(f) versus ln[(δf/1000 + 1)/(δ0/1000 + 1)], where f is the fraction remaining, δf is δ13C or δ2H at time f, and δ0 is the initial δ13C or δ2H value. The linear regression of this plot yields a slope (R - 1) which is related to the enrichment factor (ε) such that ε ) 1000(R - 1) (18). The 95% confidence intervals (CIs) for the regressions were also calculated. To VOL. 43, NO. 8, 2009 / ENVIRONMENTAL SCIENCE & TECHNOLOGY
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TABLE 2. Carbon Isotopic Values (δ13C) of Parent Fuel Oxygenates and Daughter Products (TBA and TAA) Produced during Biodegradation by Pseudonocardia K1 and Abiotic Acid Hydrolysisa fuel oxygenate MTBE ETBE TAME
raw materials used to synthesize the fuel oxygenate (1)
degradation method
initial δ13C of parent (‰)
δ13C (‰) of tertiary alcohol daughter product
calcd δ13C (‰) of released groupc
isobutene and methanol isobutene and ethanol isoamylene and methanol
biodegradation acid hydrolysis biodegradation acid hydrolysis biodegradation acid hydrolysis
-30.7 -30.9 -22.3 -24.7b -23.8 -23.8
TBA, -28.2 TBA, -29.9 (98.7% reacted) TBA, -28.4 TBA, -30.3 (99.2% reacted) TAA, -22.4 TAA, -19.8 (99.9% reacted)
methoxy, -40.7 methoxy, -34.9 ethoxy, -10.1 ethoxy, -13.5 methoxy, -30.8 methoxy, -43.8
a The δ13C of released methoxy and ethoxy groups is calculated on the basis of mass balance after Zwank et al. (8). The total analytical uncertainty for each δ13C value is ( 0.5‰ incorporating both accuracy and reproducibility (13, 31). b The initial δ13C of ETBE used is different from that in the biodegradation study. c δ13CMTBE ) (4/5)δ13CTBA + (1/5)δ13Cmethoxy, δ13CETBE ) (4/6)δ13CTBA + (2/6)δ13Cethoxy, and δ13CTAME ) (5/6)δ13CTAA + (1/6)δ13Cmethoxy.
FIGURE 2. Carbon isotopic values versus the fraction of parent compound remaining during aerobic biodegradation of (a) MTBE, (b) ETBE, and (c) TAME by Pseudonocardia K1. In all cases, results from all repeat experiments are plotted. The parent compounds are represented by closed symbols. The solid line represents the Rayleigh fractionation curve for the parent products calculated using all data (Table 1). The open symbols are the daughter product (TBA or TAA) produced during the experiment. The dashed lines are the expected δ13C value of TBA or TAA produced from oxidation (red) or SN1 (blue) reaction mechanisms. Vertical error bars represent the total uncertainty (accuracy and reproducibility) of (0.5‰ on δ13C values (13, 31). calculate AKIEs, the position-specific enrichment factor (εreactive position) was used. εreactive position takes into account the nonreacting positions within the compound and is calculated by plotting ln(f) versus ln[(1000 + δ0 + n/x(δf - δ0))/(1000 + δ0)], where n is the number of atoms and x is the number of atoms in reacting positions (7, 20). Intermolecular competition was corrected according to AKIE ) 1/[1 + z*εreactive position/1000]
(3)
where z is the number of chemically equivalent positions that are in intramolecular competition (7, 20). For carbon AKIE calculations, x ) 1 and z ) 1 since there is only one carbon at the reaction site. For hydrogen AKIE calculations for the oxidative mechanism, x and z are both equal to 3 since there are three hydrogen at the reaction site in intramolecular competition. For the SN1 and SN2 mechanisms, z ) 1 since the hydrogen atoms do not participate directly in the initial reaction, and the value of x pertains to the number of atoms located close to the reacting bond in the MTBE molecule (i.e., x ) 9 for SN1 and x ) 3 for SN2). The correction factors n, x, and z for MTBE, ETBE, and TAME for the possible biodegradation mechanisms are reported in Figure 1 and Table SI-1 in the Supporting Information.
Results and Discussion Carbon and Hydrogen Isotopic Fractionation of Parent Compounds. The initial δ13C of MTBE was -30.7 ( 0.5‰ (Table 2). Significant enrichment in 13C occurred during MTBE biodegradation, corresponding to an MTBE εC of -2.3 ( 0.2‰ (r2 ) 0.98) (Figure 2, Table 1). Data for individual bottles in all experiments are presented in Table SI-2 in the Supporting Information. The initial δ13C values of ETBE and TAME were -22.3 ( 0.5‰ and -23.8 ( 0.5‰, respectively 2796
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(Table 2). The magnitudes of εC for ETBE (-1.7 ( 0.2‰, r2 ) 0.98) and TAME (-1.7 ( 0.3‰, r2 ) 0.99) were smaller than for MTBE (Figure 2, Table 1), consistent with the fact that, compared to MTBE, both ETBE and TAME have an additional carbon atom which dilutes the measured isotopic fractionation (Table SI-1). After correction for the different numbers of carbon atoms (n), AKIEs for the oxidation of all three ethers by strain K1 were consistently 1.01 (Table 1), supporting the idea that biodegradation occurs via a similar mechanism for all three compounds. The carbon AKIE for biodegradation of MTBE, ETBE, and TAME by strain K1 of 1.01 is smaller than the expected range for an SN2 hydrolysis reaction, suggesting this mechanism does not occur. However, on the basis of carbon AKIEs alone, it is difficult to distinguish between the C-H oxidation reaction and an SN1 hydrolysis reaction since the measured carbon AKIE for K1 of 1.01 fits with the expected range for both of these mechanisms (Table 1, Figure 1). The initial δ2H value of MTBE was -77‰. Significant enrichment in 2H occurred during MTBE biodegradation, corresponding to an MTBE εH of -100 ( 10‰ (r2 ) 0.99, Table 1). The δ2H values of ETBE and TAME at the start of the experiments were -129‰ and -92‰, respectively. The magnitude of εH for ETBE (-73 ( 7‰, r2 ) 0.97) and TAME (-72 ( 2‰, r2 ) 0.99) is smaller than for MTBE (-100 ( 10‰) during aerobic biodegradation by strain K1 (Table 1). Again, this is consistent with the fact that both ETBE and TAME have two additional hydrogen atoms compared to MTBE, which results in a dilution of the measured isotopic fractionation (Table SI-1, Supporting Information). Hydrogen AKIEs were calculated in an attempt to discern between the two remaining proposed mechanisms. The calculated hydrogen AKIE values for MTBE, ETBE, and TAME biodegradation by K1 assuming the oxidative and SN1 hydrolysis
FIGURE 3. (a) Changes in carbon and hydrogen isotope ratios during aerobic biodegradation of fuel oxygenates by Pseudonocardia K1 (b, MTBE, ETBE, and TAME) and M. petroleiphilum PM1 ([, MTBE) (13). The solid lines represent the regression of the data for K1 and PM1, and the dashed lines are the 95% confidence intervals. The two VAFB microcosm replicates for MTBE biodegradation are shown as O and ] (13). Propagated errors for CSIA are shown ((0.7‰ for ∆13C, (7‰ for ∆2H) (13, 31). (b) Enlarged version of the inset from (a). mechanisms are presented in Table 1. The observed hydrogen AKIEs for all three oxygenates fit with the expected ranges for both oxidation (2-50) and SN1 hydrolysis (1.1-1.2) mechanisms (Figure 1, Table 1). Assuming an oxidative mechanism of ether degradation by strain K1, the hydrogen AKIEs for MTBE, ETBE, and TAME were estimated to be 14.9, 4.2, and 4.5, respectively. Plots comparing ∆2H (δ2Hf - δ2H0) versus ∆13C (δ13Cf - δ13C0) for MTBE, ETBE, and TAME biodegradation (Figure 3) show that the slopes of ∆2H/∆13C are the same (Table 1), suggesting all three oxygenates are degraded by the same biodegradation mechanism. Isotopic fractionation of daughter products was measured to further distinguish between the oxidation and SN1 mechanisms. δ13C Measurements of TBA and TAA. Since Pseudonocardia strain K1 does not further oxidize TBA and only slowly oxidizes TAA in the absence of TAME, δ13C measurements of these biodegradation “end products” can provide important additional information about the prevailing biodegradation mechanism (5, 6). For both the oxidation and SN2 mechanism, the carbon atoms on the tert-butyl or tert-amyl group do not participate in the initial reaction (Figure 1) and the carbon atoms in the produced tertiary alcohol would not be subject to significant isotopic fractionation. For SN1 reactions where the O bond to the tert-butyl or tert-amyl group is broken, the preferential release of tertiary groups containing 12C would be expected to result in instantaneously formed TBA depleted in 13C compared to the tert-butyl group from which it was formed. Consequently, for the SN1 mechanism both TBA and TAA would show a trend of isotopic enrichment over time until all of the ether has been biodegraded and the final δ13C value of the tertiary alcohol reflects the δ13C of the original tertiary groups in the substrate due to quantitative (100%) conversion. The δ13C of TBA produced during aerobic biodegradation of MTBE by strain K1 was constant through time with a mean of -28.2 ( 0.4‰ (1σ std dev). This δ13C TBA value is slightly enriched compared to that of the parent MTBE (δ13C of -30.7‰) (Table 2, Figure 2a), in agreement with an aerobic biodegradation study by Hunkeler et al. (14). During ETBE and TAME experiments, the produced TBA (δ13C of -28.4 ( 0.4‰, 1σ std dev) and TAA (δ13C of -22.4 ( 0.2‰, 1σ std dev) were also constant throughout the experiments (Table 2, Figure 2b,c; for data of all individual replicate bottles, see Table SI-2 in the Supporting Information). The invariant TBA and TAA values through time suggest that the carbon atoms on the tert-butyl or tert-amyl groups do not participate in the reaction and strongly suggest oxidation as the prevailing mechanism for biodegradation of all three fuel oxygenates by K1. To further illustrate this point, the expected δ13C TBA and TAA values for the SN1 mechanism were calculated (as
described in the Supporting Information) and added to Figure 2. The δ13C values of measured TBA and TAA fit with the expected values for oxidation rather than the values expected for the SN1 enzymatic mechanism. Acid hydrolysis of the ethers to TBA or TAA daughter products was also conducted, after Zwank et al. (7), to confirm the observed enrichment (Figure 2a,c) or depletion (Figure 2b) in 13C of the daughter tertiary alcohols relative to the parent compounds. Following abiotic acid hydrolysis, the TBA daughter product of MTBE is indeed enriched in 13C compared to the initial MTBE (Table 2). Similarly, TAA is enriched in 13C relative to parent TAME. In contrast, TBA produced from acid hydrolysis of ETBE is depleted in 13C, compared to the initial ETBE (Table 2). This is because the ethoxy group for ETBE is isotopically enriched in 13C compared to the methoxy groups for MTBE and TAME, reflecting that it originates from ethanol of agricultural origin rather than petroleum-based methanol (1). Comparison with Other Strains. The hydrogen isotopic enrichment factor (εH) for MTBE biodegradation by strain K1 of -100 ( 10‰ is the largest reported to date for MTBE biodegradation and agrees well with the measured εH for abiotic oxidation of MTBE by permanganate, which had an εH value of -109 ( 9‰ (21). Elsner et al. (21) suggested that in microbial systems where the breakage of the C-H bond is rate determining, the εH measured during biodegradation could approach the εH value of permanganate oxidation. The agreement between εH for abiotic and enzymatic reactions suggests that the isotopic effect is mainly associated with the breakage of the C-H bond in strain K1. Alternately, if additional rate-limiting reaction steps occur, the measured εH would be lower (21, 22). On this basis, it could be argued that, during biodegradation of MTBE by strain K1, the ratelimiting step is breakage of the C-H bond, and Methylibium strains PM1 (13) and R8 (15) show smaller hydrogen isotopic fractionation due to additional rate-limiting steps (7, 21). To determine whether the differences in hydrogen isotopic fractionation between strains K1 and PM1 were due to additional rate-limiting steps or potentially due to differences in biodegradation mechanisms, plots of ∆2H versus ∆13C were compared (Figure 3). During biodegradation of MTBE, different slopes for ∆2H/ 13 ∆ C for strain K1 (∆2H/∆13C ) 48) compared to Methylibium strain PM1 (∆2H/∆13C ) 18 (13)), suggest that different initial mechanisms may be utilized during biodegradation by these organisms (Figure 3). There is strong isotopic evidence supporting an oxidative mechanism for MTBE biodegradation by Pseudonocardia strain K1, including (a) carbon and hydrogen AKIE values for K1 consistent with the oxidative mechanism, (b) δ13C measurements of the daughter TBA VOL. 43, NO. 8, 2009 / ENVIRONMENTAL SCIENCE & TECHNOLOGY
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which show that the carbon atoms in the tert-butyl group do not participate in the reaction, and (c) consistency of K1 εH values with the εH values for abiotic oxidation. While it can be inferred that PM1 and R8 utilize the same biodegradation mechanism on the basis of their similar slopes (15), the present study questions earlier interpretations which attributed the ∆2H/∆13C of PM1 and R8 to a C-H bond oxidation mechanism (7, 15, 21), since it is shown here that the slope of ∆2H/∆13C for oxidation is considerably higher (slope of 48 ( 5) than previously thought. Neither does the isotopic fractionation exhibited by strains PM1 and R8 appear consistent with the SN1 hydrolysis mechanism for which a ∆2H/∆13C slope of 11 ( 1 has been measured during abiotic degradation (21). Unfortunately, since strains PM1 and R8 do not produce measurable TBA (13, 15, 32), isotopic measurement of this daughter product, which effectively demonstrated that K1 utilizes the oxidative mechanism, could not be applied to PM1 and R8. Additional isotopic, enzymatic, and microbial investigations are necessary to elucidate the mechanism by which these two strains biodegrade MTBE. One possibility is the measurement of δ18O during biodegradation of MTBE, which could provide valuable insight into the biodegradation mechanisms used by these microorganisms since the initial breakage of the ether bond would likely result in measurable 18O fractionation, whereas the breakage of a C-H bond on the methoxy group would not. Gray et al. (13) found variability in hydrogen isotopic fractionation between replicate bottles in a laboratory study investigating biodegradation of MTBE in microcosms from VAFB (εH of -29‰ and -66‰). While carbon εC values also differed for the replicates, they showed a much smaller range (εC of -1.4‰ to -1.8‰). It was hypothesized that the variability might be due to differences in the microbial community that had arisen through time in the different replicate bottles. The ∆2H/∆13C for the two replicates has been added to Figure 3. The data points for the two microcosms fall within or close to the confidence intervals for strains K1 and PM1, respectively, suggesting that the variability in δ2H values in the earlier study might be due to differences in the microbial population in the replicate bottles over time. To date, there have been two separate trends in isotopic enrichment factors shown for biodegradation of MTBE under aerobic conditions (Table 1, Figure 3), suggesting the possibility of two different biodegradation mechanisms.
Acknowledgments Funding for this project was provided by the Natural Sciences and Engineering Research Council of Canada and the American Petroleum Institute (API). Scholarships provided to J.R.M. by the API, National Ground Water Association, and Government of Ontario are acknowledged.
Supporting Information Available (SI1) experimental details, (Table SI-1) parameters for calculation of AKIEs, (Table SI-2) isotopic data for the individual replicates, and (SI2) calculations of the expected δ13C values of TBA and TAA produced via an SN1 mechanism. This material is available free of charge via the Internet at http://pubs.acs.org.
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