Changes in Enantiomeric Fractions during Microbial Reductive

Feb 7, 2003 - Mazur and Jones (21). Sediment slurries from only the G30 location were prepared inside an anaerobic chamber con- taining an atmosphere ...
0 downloads 0 Views 186KB Size
Download PDF
Environ. Sci. Technol. 2003, 37, 1100-1107

Changes in Enantiomeric Fractions during Microbial Reductive Dechlorination of PCB132, PCB149, and Aroclor 1254 in Lake Hartwell Sediment Microcosms USARAT PAKDEESUSUK,† W . J A C K J O N E S , ‡ C I N D Y M . L E E , * ,† ARTHUR W. GARRISON,‡ WALTER L. O’NIELL,‡ DAVID L. FREEDMAN,† JOHN T. COATES,† AND C H A R L E S S . W O N G §,⊥ Department of Environmental Engineering & Science, Clemson University, Clemson, South Carolina 29634-0919, National Exposure Research Laboratory, Ecosystems Research Division, U.S. Environmental Protection Agency, Athens, Georgia 30605-2700, and Department of Chemistry, University of Toronto, Toronto, Ontario M5S 3H6, Canada

The enantioselectivity of microbial reductive dechlorination of chiral PCBs in sediments from Lake Hartwell, SC, was determined by microcosm studies and enantiomer-specific GC analysis. Sediments from two locations in the vicinity of the highest levels of PCB contamination were used as inocula. Dechlorination activity was monitored by concentration decreases in the spiked chiral PCBs and formation of dechlorination products using both achiral and chiral chromatography. Live microcosms spiked with PCB132 (234-236) exhibited dechlorination of PCB132 to PCB91 (236-24) and PCB51 (24-26). Meta dechlorination was the dominant mechanism. Microcosms spiked with PCB149 (245-236) exhibited preferential para dechlorination of PCB149 to PCB95 (236-25), followed by meta dechlorination to PCB53 (25-26) and subsequently PCB19 (26-2). Dechlorination of chiral PCB132 and PCB149 was not enantioselective. In Aroclor 1254-spiked microcosms, reductive dechlorination of PCB149 also was nonenantioselective. These results suggest that dechlorinating enzymes responsible for the dehalogenation of the chiral PCB132 and PCB149 congeners bind the two enantiomers equally. Reductive dechlorination of PCB91 and PCB95, however, occurred in an enantioselective manner, indicating that the dechlorinating enzymes for these PCBs are enantiomer-specific. The chlorine substitution pattern on the biphenyl ring appears to influence whether reductive dechlorination of chiral PCB congeners is enantioselective. Enantioselective PCB dechlorination by the microbial population of Lake Hartwell sediments occurs for select chiral PCBs; thus, certain chiral PCBs might be useful as markers for in situ reductive dechlorination. * Corresponding author phone: (864) 656-1006; fax: (864) 6560672; e-mail: [email protected]. † Clemson University ‡ U.S. Environmental Protection Agency. § University of Toronto ⊥ Present address: Department of Chemistry, University of Alberta, Edmonton, AB T6G 2G2, Canada. 1100

9

ENVIRONMENTAL SCIENCE & TECHNOLOGY / VOL. 37, NO. 6, 2003

These results represent the first evidence of stereoselective reductive dechlorination of PCBs under controlled conditions.

Introduction Microbial reductive dechlorination is the major degradation mechanism for PCBs under anaerobic conditions (1, 2). In situ reductive dechlorination of PCBs has been inferred from congener-specific analyses based on the correlation of losses in more chlorinated PCB congeners and increases in less chlorinated congeners that are considered the residual products (1-4). To provide support for in situ reductive dechlorination as inferred from changes in field congener profiles, the presence and activity of PCB-dechlorinating microorganisms should be confirmed by, for example, a microcosm study. However, without chronological field data, it is often difficult to provide compelling proof that biologically mediated processes are resulting in natural attenuation at a given site solely on the basis of the disappearance of parent compounds and the accumulation of final products. This is especially true for a site such as Lake Hartwell, SC, where three different Aroclors were used by the manufacturer responsible for the contamination and where volatilization and sediment transport have affected the distribution of congeners (5). Analysis of enantiomers (nonsuperimposable mirror images of a chiral compound) provides a promising approach to elucidate the importance or contribution of biological activities, particularly microbial degradation, in a given system (6-8). Whereas the chemical and physical properties of the enantiomers of chiral compounds are identical, enantiomers generally have different biological properties, especially the rate at which they interact with other chiral compounds, such as enzymes and other biochemicals (6). Therefore, interaction of the enantiomers of a chiral compound within a biological system often results in enantioselectivity, in which one enantiomer is degraded faster than the other. For example, (S)-mecoprop is degraded faster than (R)-mecoprop by Spingomonas herbicidovoran MH (9), whereas the reverse enantioselectivity occurs with Alcaligenes denitrificans (10). On the other hand, Garrison et al. (6) did not observe enantioselective degradation of o,p′-DDT in the presence of several plant species or in anaerobic sediments, although enantioselectivity was reported by Wiberg et al. (11) for o,p′-DDT in soil from different locations. Some researchers have suggested that the shift in enantiomeric fraction (EF) of a chiral compound might be a useful marker for microbially mediated processes (7). The EF is the fraction of the concentration of the (+)-enantiomer over the sum of the concentration of both enantiomers or, when the elution order is unknown, the EF is the concentration of the first eluted enantiomer over the concentration sum for both (12). However, it should be noted that not all enzymes are enantioselective toward a particular chiral compound and more must be discovered about the specificity of enzymatic reactions in environmental systems. Lewis et al. (13) showed that different environmental conditions can affect the enantioselectivity of microbial degradation of chiral contaminants by favoring specific microbial populations of related genotypes. Atropisomeric PCBs are considered an important class of chiral organic pollutants. Nineteen of the possible 209 PCB congeners contain three or four o-chlorine atoms and exist as pairs of stable enantiomers; these congeners display axial chirality in their nonplanar conformation and encounter energy barriers to rotation about the central C-C bond greater 10.1021/es026039g CCC: $25.00

 2003 American Chemical Society Published on Web 02/07/2003

than 83 kJ mol-1 (14, 15). Enantioselective reductive dechlorination can affect the enantiomeric composition of chiral PCBs in sediments, yielding a deviation of the EF from the value of racemates (EF ) 0.5). Recently, the enantiomeric composition of chiral PCBs in the sediments of Lake Hartwell was reported and proposed as a potential indicator of microbial PCB dechlorination (7). Whereas the enantiomers of some PCB chiral congeners were present at nonracemic quantities in sediments from several locations in Lake Hartwell, a lack of enantioselectivity was also observed at other locations. This lack of enantioselectivity might be due to the absence of PCB dechlorination or to nonenantioselective reductive dechlorination. Although there are some reports of marked changes in the enantiomeric composition of chiral PCBs in sediments (7, 16, 17), it has not yet been shown whether reductive dechlorination of PCB enantiomers is the enantioselective process responsible for the preferential degradation of one enantiomer over the other. Further information on enzymatic systems that lead to enantioselective dechlorination pathways of chiral PCBs is needed, as well as on the mechanisms responsible for the observed enantioselectivity. The goals of this study were to determine whether the microbial communities in sediments of Lake Hartwell were capable of reductive dechlorination of chiral PCBs and whether reductive dechlorination of the PCB enantiomers and their chiral products was enantioselective. Advancing our understanding of enantioselective processes should help in elucidating the mechanisms for the transformation of chiral chemicals such as PCBs in natural ecosystems.

Materials and Methods Sediment Collection. The sediments used in this study were collected from the Twelve Mile Creek arm of Lake Hartwell, which is the most heavily PCB-contaminated location in the lake (5). The sampling locations (G30 and G33) are shown in Figure 1. Sediment cores were collected in July 1998 using a Wildco sediment corer containing a Lexan tube (5-cm diameter, 76-cm length) and 5.1-cm eggshell core catcher. The cores were transported to the laboratory and stored in the dark prior to sectioning and subsampling. The top layer (about 5 cm) of each core was removed, and the remaining portion was transferred to a glass jar, topped with site water, sealed, and stored at 4 °C prior to use. Microcosms. One set of microcosms was prepared for evaluating 234-236 hexachlorobiphenyl (IUPAC number PCB132) and Aroclor 1254, and a second set was prepared for 245-236 hexachlorobiphenyl (PCB149). Triplicate serum bottles were used for each treatment. The PCB132 and Aroclor 1254 microcosms were set up as described by Pakdeesusuk (18). Sediment slurries from the G30 and G33 sites were prepared inside an anaerobic chamber containing an atmosphere of 98% N2 and 2% H2. The collected sediment was homogenized and thoroughly mixed with sulfide-reduced mineral media (19; NaHCO3 increased to 4.0 g/L) to achieve a sediment solids concentration of approximately 100 g per liter of slurry. Aliquots (50 mL) of the sediment slurry were dispensed into amber serum bottles that were subsequently purged with a nitrogen/carbon dioxide (70%/30%) gas mixture and sealed with Teflon-lined butyl rubber septa and aluminum crimp caps. The final pH ranged from 6.5 to 7.0. A concentrated stock solution of PCB132 or Aroclor 1254 (AccuStandard Inc.) dissolved in acetone was added to a final concentration of approximately 500 µg/g. To ensure a homogeneous distribution of PCBs, the serum bottles were shaken overnight. They were then incubated in an upright position and in the dark without shaking in the anaerobic chamber at room temperature (2224 °C). Killed controls were prepared identically except that the bottles containing sediment slurry were autoclaved (121

°C, 15 min) for three consecutive days prior to addition of the PCBs. Samples of sediment slurry were collected in the anaerobic chamber at various time points to analyze for acetone and organic acids after headspace analysis for CH4 and H2 using the analytical methods described elsewhere (18, 20). For PCB analyses, subsamples (3-5 g) were removed and extracted by sonication in acetone (5). After sampling, the serum bottles were purged with the N2/CO2 gas mixture, resealed, and returned to the anaerobic chamber for continued incubation. The PCB149 microcosms were set up as described by Mazur and Jones (21). Sediment slurries from only the G30 location were prepared inside an anaerobic chamber containing an atmosphere of 98% N2 and 2% H2. The collected sediment was homogenized and thoroughly mixed with anoxic (N2-sparged) site water to achieve a sediment solids concentration of approximately 100 g/L of slurry. PCB149 (dissolved in acetone) was added to sterile, empty amber serum bottles to achieve a final concentration of 10 µg/g in the slurry, and the solvent was evaporated to dryness. Aliquots (5 mL) of the sediment slurry were then dispensed into the bottles containing PCB149, which were subsequently sealed with Teflon-lined butyl rubber septa and aluminum crimp caps. The pH of the slurry was 6.5. To ensure homogeneous distribution of PCB149, the serum bottles were vigorously shaken for 24 h. The bottles were incubated statically in an upright position and in the dark at 22 °C. Killed controls were prepared by autoclaving (121 °C, 15 min) the sediment slurry for three consecutive days, followed by aseptic addition of the slurry (5 mL) to serum bottles containing the solventevaporated PCB149 solution. Replicate experimental and control serum bottles were prepared to allow for complete sacrifice of the microcosms at specified sampling times. PCB extraction from the sediment slurry microcosms was performed according to the method of Quensen et al. (22). Achiral Chromatography. PCB149, PCB132, and dechlorination products were identified by matching GC retention times with individual PCB congeners (AccuStandard Inc., New Haven, CT). PCBs were initially quantified using a gas chromatograph (Hewlett-Packard 6890) equipped with an achiral column (30 m × 0.25 mm × 0.25 µm, ZB5, Phenomenex), and a 63Ni electron-capture detector (ECD) according to the method of Dunnivant and Elzerman (23) as modified by Farley et al. (5). Helium (2 mL/min) was used as a carrier gas, with the same temperature program and instrument conditions as previously described (23), except that the initial hold time was changed to 2.5 min. Products of the dechlorination of the two chiral hexachlorobiphenyls PCB132 and PCB149 were predicted by assuming that only meta or para dechlorination occurred. Previous work indicated that ortho chlorines were resistant to reductive dechlorination by Lake Hartwell sediments (18, 20). For the PCB132 (234-236) product analysis, the following standards were purchased from AccuStandard, Inc.: PCB2 (2), PCB4 (2-2), PCB17 (24-2), PCB19 (26-2), PCB45 (236-2), PCB46 (23-26), PCB51 (24-26), PCB84 (236-23), PCB89 (234-26), and PCB91 (236-24). Identification was accomplished by matching GC retention times. Four-point calibration curves for identified products were built using aldrin (AccuStandard, Inc.) as an internal standard. PCB congeners in Aroclor 1254 were identified and quantified using the methods described previously (5, 18, 20, 23). Chiral Chromatography. Extracts from PCB149- and PCB132-spiked microcosms were analyzed for chiral PCB149, PCB132, PCB95, and PCB91 by a capillary gas chromatograph (Hewlett-Packard 5890) equipped with a 63Ni electron-capture detector (GC/ECD) (24). The instrument was equipped with a Chirasil-Dex column (25 m × 0.25 mm × 0.25 µm, Chrompack). A 1-µL aliquot was injected in the split mode (ratio 1:65). Helium (2 mL/min) was used as the carrier gas. VOL. 37, NO. 6, 2003 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

9

1101

FIGURE 1. Location of Lake Hartwell, the Twelve Mile Creek arm of the Lake, and sampling points for this study: G30 latitude 34°43′12′′, longitude 82°49′61′′; G33 latitude 34°42′30′′, longitude 82°50′50′′. The temperatures of injector and detector were maintained at 250 and 325 °C, respectively. The column temperature was programmed from 100 to 150 °C at 10 °C/min and then from 150 to 200 °C at 0.5 °C/min. Extracts from the Aroclor 1254-spiked microcosms were analyzed for chiral PCBs 91, 95, 136, and 149 using a CycloSil-B (J&W Scientific) chiral column and by using the chiral 1102

9

ENVIRONMENTAL SCIENCE & TECHNOLOGY / VOL. 37, NO. 6, 2003

analytical procedure described previously (25). Note that PCB132 could not be separated satisfactorily in samples from the Aroclor 1254-spiked microcosms. PCB136 (236-236) was also analyzed, although it cannot serve as a surrogate for PCB132. Enantiomeric fractions (EFs) were calculated according to eq 1 (12)

EF ) area of (+)-enantiomer/[area of (+)-enantiomer plus area of (-)-enantiomer] (1) EFs for PCBs 91 and 95 were expressed as area ratios of the first-eluting peaks to the sum of the two peaks on the ChirasilDex column (7), as the enantiomer elution order is unknown (no enantiomer standards were available). Replicate injections of analytical standards of all measured enantiomers produced the same enantiomer peak areas within experimental error, in accord with racemic compositions (EF ) 0.500 ( 0.005).

Results and Discussion Dechlorination and Enantioselective Transformation of Selected Chiral PCBs. Microcosm results with sediment from the G30 and G33 sites of the Twelve Mile Creek arm of Lake Hartwell confirmed the presence of indigenous microbial populations that are capable of reductive dechlorination of the chiral PCB congeners 234-236 CB (PCB132) and 245-236 (PCB149). Decreases in the PCB132 and PCB149 concentrations occurred in live treatments but not in autoclaved controls, indicating that the dechlorination processes were biologically mediated (Figures 2A and 3A). PCB mass balances for both PCB132- and PCB149-spiked sediment microcosms were calculated for autoclaved controls and live treatments, with good recoveries of 80.6% ((11.0) and 102.1% ((15.4), respectively, at the final incubation times. In the G30 live microcosms, the decrease in PCB132 (Figure 2A) was accompanied by the formation of two major dechlorination products: 236-24 CB (PCB91) and 24-26 CB (PCB51) (Figure 2B,C). The likely dehalogenation pathway is shown in Figure 4A. PCB91 and PCB51 were detected almost simultaneously (Figure 2B,C), suggesting that dechlorination of PCB91 to PCB51 occurred at a faster rate than dechlorination of PCB132 to PCB91. A dechlorination pathway to PCB51 via 234-26 CB (PCB89) appeared less likely, because PCB89 was not detected. Formation of PCB91 and PCB51 from PCB132 indicated a preference for meta dechlorination. This same preference was observed in Lake Hartwell G30 and G33 microcosms spiked with Aroclor 1254 (18, 20). Near the end of the incubation period, a low rate of para dechlorination of PCB51 yielded small amounts of 26-2 CB (PCB19) (data not shown). At no time were any products detected that were indicative of ortho dechlorination. In the G33 live microcosms, only one of the triplicate slurries exhibited a significant level of PCB132 dechlorination activity (data not shown). The pathway of PCB132 dehalogenation for this G33 microcosm was similar to that for the G30 microcosms, resulting in the accumulation of PCB91 and PCB51. The reason for a lack of dehalogenation activity in the replicate microcosms is unknown but might be an initially low level of PCB132-dehalogenating microbes in G33 sediments such that experimental replication was variable. Kim and Rhee (26) observed this phenomenon when applying the most probable number approach to quantifying PCBdechlorinating microorganisms. In other microcosms prepared with G33 sediment, dechlorination of Aroclor 1254 occurred in all three of the live microcosms tested (18, 20). Dechlorination of PCB149 was observed in G30 live sediment microcosms after a lag period of approximately 65 days (Figure 3A). The preferential para dechlorination of PCB149 resulted in the formation of 236-25 (PCB95), followed by sequential meta dechlorination to produce 25-26 (PCB53) and 26-2 (PCB19). The proposed dehalogenation pathway is shown in Figure 4B. PCB53 and PCB19 were not detected until after the formation of PCB95, suggesting that the dechlorination of PCB149 to PCB95 was faster than dechlorination of either PCB95 to PCB53 or PCB53 to PCB19 (Figure 3). Formation of PCB91 was not observed, eliminating the likelihood of an initial meta dechlorination pathway of

FIGURE 2. Concentration of PCB132 and its major dechlorination products (A-C) and enantiomeric fractions (EFs) for PCB132 and PCB91 (D, E) in PCB132-spiked G30 microcosms with time: (open circles, crosshatched bars) autoclaved controls with PCB132, (filled circles, filled bars) live treatments with PCB132, (- - -) EF ) 0.5. Error bars represent standard deviations for triplicate bottles; where not shown, the deviation was smaller than the size of the symbols. PCB149. No dechlorination products were detected that indicated an ortho dehalogenation reaction. PCB132 and PCB149 were selected for use in this study because both are chiral PCBs and are significant constituents of commercial Aroclors 1254 and 1260. However, no evidence of enantioselective reductive dechlorination of PCB132 and PCB149 was detected in the live, PCB-spiked G30 or G33 microcosms. Constant racemic EF values of approximately 0.5 were observed for PCB132 and PCB149 enantiomers throughout the incubation periods (Figures 2D and 3D). There were also no changes in EF values in the autoclaved controls. However, dechlorination of chiral product PCB91 (resulting from the meta dechlorination of PCB132), as well as dechlorination of chiral product PCB95 (resulting from para dechlorination of PCB149), was enantioselective (Figures 2E and 3E). At the first detection of PCB91in the live treatments, VOL. 37, NO. 6, 2003 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

9

1103

FIGURE 3. Concentration of PCB149 and its major dechlorination products (A-C) and enantiomeric fractions (EFs) for PCB149 and PCB95 (D, E) in PCB149-spiked G30 microcosms with time: (open circles, crosshatched bars) autoclaved controls with PCB149, (filled circles, filled bars) live treatments with PCB149, (- - -) EF ) 0.5. Error bars represent standard deviations for triplicate bottles; where not shown, the deviation was smaller than the size of the symbols. its EF values were 0.42 in the G30 microcosm (Figure 2E) and 0.26 in the G33 microcosms that exhibited activity (data not shown). An observed EF of less than 0.5 signifies that the rate of dechlorination for the first-eluted enantiomer (E1) was higher than that for the other enantiomer. During further incubation (up to 166 days), the EF of PCB91 in the G30 microsoms and single G33 microcosm shifted to even lower values (), minor pathway. sediments. The live microcosms actively dechlorinated Aroclor 1254 by a pathway similar to pattern M (1, 18, 20), characterized by preferential removal of unflanked and flanked m-chlorines. In the present study, we measured the EF values of four chiral PCB congeners (PCBs 149, 136, 95, and 91) from the Aroclor 1254-spiked microcosms (Table 1). Note that the EF for PCB132 was not determined in the Aroclor 1254-spiked microcosms because of coeluting interferences from hexachloro congeners in the Aroclor mixture. However, PCB136 (236-236) was successfully separated by a chiral column (CycloSil-B) and quantified. PCB136 is not considered a surrogate for PCB132 (234-236). The behavior of the two pentachlorobiphenyls, PCBs 91 (236-24) and 95 (236-25), in the Aroclor 1254-spiked microcosms supported the results observed in the microcosms spiked with the single congeners. EF values provided clear evidence for enantioselective transformation of PCB91, on the basis of decreases in EF values for PCB91 after 2-3 months of incubation, in both G30 and G33 microcosms. With continued incubation, the EF for PCB91 decreased further,

pointing to a dechlorinating enzyme for PCB91 that was enantiomer-specific. Supporting the decrease in EF, achiral analyses revealed a reduction in the weight percent of the peaks containing PCB91 (236-24) and its coeluted congeners, 246-35 CB and 234-3 CB, in the Aroclor 1254-spiked microcosms (18, 20). However, the possibility exists that enantioselective formation of PCB91 from PCB132 or some other congener occurred in the Aroclor 1254-spiked microcosm because we were unable to determine whether the disappearance of PCB132 in the mixture was racemic as was observed in the single-congener microcosm. Reductive dechlorination of PCB95 also occurred in an enantioselective manner in the G30 and G33 microcosms spiked with Aroclor 1254, although the change in EF was not as large as it was for PCB91 (Table 1). The EF for PCB95 increased above the racemic value of 0.5 after 2-3 months of incubation, in the opposite direction to the changes in the EF of PCB91. However, because the elution order of the (()enantiomers for both PCB91 and PCB95 was unknown, it is not possible at this time to draw any conclusions regarding VOL. 37, NO. 6, 2003 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

9

1105

FIGURE 5. Comparison of (A,B) methane formation, (C,D) acetate concentration, and (E,F) acetone concentration in G30 and G33 microcosms: (open circles) autoclaved controls spiked with PCB132 in acetone carrier solvent, (filled circles) live treatments spiked with PCB132 in acetone carrier solvent, (filled triangles) live treatments without acetone or PCB132, (I) lag phase prior to PCB dechlorination, (II) rapid phase of dechlorination, (III) plateau phase of dechlorination. Error bars represent standard deviations for triplicate bottles; where not shown, the deviations were smaller than the size of the symbols.

TABLE 1. EF Values of Four Chiral PCBs in Aroclor 1254-Spiked Microcosms PCB91

PCB136

PCB149

control

live

control

live

control

live

control

live

G30

0 50 77 99 112 140

0.51 0.51 0.51 0.51 0.51 0.51

0.51 0.51 0.50a 0.35 0.14 0.14a

0.49 0.49 0.49 0.49 0.49 0.49

0.49 0.49 0.50a 0.57 0.54 0.58a

0.50 0.51 0.51 0.50 0.51 0.51

0.51 0.50 0.51a 0.50 0.60 0.55a

0.49 0.48 0.49 0.49 0.49 0.49

0.49 0.48 0.49a 0.43 0.51 0.49a

G33

0 26 67 112 136 166 258 354

0.51 0.50 0.50 0.50 0.50 0.50 0.50 0.50

0.51 0.50 0.36 0.29 0.30 0.17 0.26 NA

0.50 0.50 0.49 0.50 0.50 0.50 0.50 0.49

0.49 0.50 0.56 0.53 0.61 0.62 NAb NA

0.51 0.51 0.50 0.51 0.50 0.51 0.51 0.50

0.50 0.51 0.50 0.51 0.50 0.56 0.51 NA

0.49 0.49 0.49 0.49 0.48 0.49 0.49 0.49

0.48 0.51 0.46 0.49 0.46 0.50 0.47 NA

site

b

PCB95

time (days)

a Average of two or three replicates; for G30, day 77, the standard deviation ranged from (0.002 to (0.008; for day 140, from (0.025 to (0.059. Not analyzed.

which enantiomers are preferentially dechlorinated. Note that, as reported above for PCB149-spiked microcosms, enantioselective transformation of PCB95 (the initial transformation product of para dechlorination of PCB149) was detected with EF values decreasing below 0.2 after 180 days 1106

9

ENVIRONMENTAL SCIENCE & TECHNOLOGY / VOL. 37, NO. 6, 2003

of incubation. Thus, the enantioselectivity of PCB95 dehalogenation in the PCB149-spiked microcosms was opposite that observed in Aroclor 1254-spiked microcosms. A possible explanation for this latter observation is that enantioselective dehalogenation of higher chlorinated chiral PCB congeners,

present in Aroclors, might generate nonracemic chiral PCBs (i.e., PCBs 149, 132, 95, and 91) with EF values greater or less than 0.5, depending on the direction of the enantioselective reaction. Achiral PCB analyses of the Aroclor 1254-spiked microcosms revealed that there was a corresponding reduction in the weight percent distribution of the peak containing PCB95 (236-25) and its coeluted congener, 24-34 CB (17, 20). Enantiomer-specific results for PCB149 in Aroclor 1254spiked microcosms indicated that PCB149 was racemic throughout the incubation period for all live treatments (Table 1). However, the weight percent distribution of the peak containing PCB149 (245-236) and its coeluted congeners, 2345-3 CB and 245-34 CB, which are minor constituents of Aroclor 1254, decreased with time (18, 20), supporting transformation of some of these congeners. It is uncertain from this experimental treatment whether the lack of change in the EF of PCB149 in the Aroclor 1254-spiked microcosms occurred because of the nonenantioselective reductive dechlorination of PCB149 or because reductive dechlorination of PCB149 did not occur. However, as reported above, significant rates of reductive dechlorination of PCB149 were observed in PCB149-spiked G30 sediments without apparent enantioselectivity. Data from Wong et al. (7) show that PCB149 is present in racemic portions at Lake Hartwell sites 30 and 33. The enantiomer results for chiral PCB136 in Aroclor 1254spiked microcosms were not definitive for the microcosm from site G33, but were for the G30 microcosm, as shown by the elevated EFs for days 112 and 140 (Table 1). Wong et al. (7) showed marked enantioselectivity for PCB136 by direct measurements of Lake Hartwell sediment from both G30 and G33, as well as other sites in the lake. In the present study, enantioselective PCB dechlorination by the microbial population of Lake Hartwell sediments varied, depending on the structural identity of the chiral PCBs. Although the reasons for this variation are not yet clear, the substitution patterns of chlorine atoms on the biphenyl rings appear to influence whether reductive dechlorination of certain PCB congeners is enantioselective. Whereas dechlorination of PCB132 and PCB149 was not enantioselective, enantioselectivity was observed during removal of the singly flanked chlorines from PCB91 and PCB95. Reductive dechlorination of PCB91 and PCB95 proceeded by removal of a flanked mchlorine from the 236 ring. For chiral phenoxyalkanoic acid derivatives, the distribution of chlorine atoms at the ortho, meta, and/or para positions on the aromatic ring influenced the enantioselectivity as well as the biodegradation rates of the enantiomers (35). Further research is needed to elucidate how chlorine substitution patterns affect enantioselective dechlorination of PCBs. The results from this study indicate that reductive dechlorination of chiral PCBs can occur enantioselectively or nonenantioselectively, depending on the chlorine substitution pattern of the PCB congeners as well as the nature of the microbial population. Changes in EF values for chiral PCBs can be used as a confirmation tool for their biodegradation, and EF values can be useful as a gross indicator of PCB biodegradation. Chiral analysis can offer us further insight into the types of degradative enzymes and microbial communities that are responsible for the transformation of PCBs in natural environments.

Acknowledgments We gratefully acknowledge support from the U.S. Environmental Protection Agency, Ecosystems Research Division, National Exposure Research Laboratory, Athens, GA. We appreciate the efforts of Jimmy Avants of ERD in performing chiral GC analyses. We also gratefully acknowledge the Savannah River Site, SC, for use of the anaerobic chamber; James L. Meyers for assistance in collecting Lake Hartwell sediments; Timur Deniz for help with drafting of the figures;

and Peter Haglund of Umea University in Sweden for the enantiomerically enriched PCBs used to identify (+)- and (-)-enantiomers. This paper has been reviewed in accordance with the U.S. Environmental Protection Agency’s peer and administrative review policies and approved for publication. Mention of trade names or commercial products does not constitute endorsement or recommendation for use by the U.S. EPA.

Literature Cited (1) Bedard, D. L.; Quensen, J. F., III In Microbial Transformation and Degradation of Toxic Organic Chemicals; Young, L. Y., Cerniglia, C. E., Eds.; Wiley-Liss, Inc.: New York, 1995; pp 127216. (2) Wiegel, J.; Wu, Q. FEMS Microbiol. Ecol. 2000, 32, 1-15. (3) Bedard, D. L.; May, R. J. Environ. Sci. Technol. 1996, 30, 237-245. (4) Sokol, R. C.; Kwon, O.-S.; Bethoney, C. M.; Rhee, G.-Yull Environ. Sci. Technol. 1994, 28, 2054-2064. (5) Farley, K. J.; Germann, G. G.; Elzerman, A. W. In Environmental Chemistry of Lakes and Reservoirs; Baker, L., Ed.; American Chemical Society: Washington, DC, 1994. (6) Garrison, A. W.; Nzengung, V. A.; Avants, J. K.; Ellington, J. J.; Jones, W. J.; Rennels, D.; Wolfe, A. N. L. Environ. Sci. Technol. 2000, 34, 1663-1670. (7) Wong, C. S.; Garrison, A. W.; Foreman, W. T. Environ. Sci. Technol. 2001, 35, 33-39. (8) Zipper, C.; Suter, M. J. F.; Haderlein, S. B.; Gruhl, M.; Kohler, H.-P. E. Environ. Sci. Technol. 1998, 32, 2070-2076. (9) Zipper, C.; Nickel, K.; Angst, W.; Kohler, H.-P. E. Appl. Environ. Microbiol. 1996, 62, 4318-4322. (10) Tett, V. A.; Willets, A. J.; Lappin-Scott, H. M. FEMS Microbiol. Ecol. 1994, 14, 191-200. (11) Wiberg, K.; Harner, T.; Wideman, J. L.; Bidleman, T. F. Chemosphere. 2001, 45, 843-848. (12) Harner, T.; Wiberg, K.; Norstrom, R. Environ. Sci. Technol. 2000, 34, 218-220. (13) Lewis, D. L.; Garrison, A. W.; Wommack, K. E.; Whittemore, A.; Steudler, P. S.; Melillo, J. Nature 1999, 401, 898-901. (14) Kaiser, K. L. E. Environ. Pollut. 1974, 7, 93-101. (15) Harju, M. T.; Haglund, P. Fresenius J. Anal. Chem. 1999, 364, 219-223. (16) Benicka`, E.; Novakovsky, R.; Hrouzek, J.; Krupcı`k, J. J. High Resolut. Chromatogr. 1996, 19, 95-98. (17) Glausch, A.; Blanch, G. P.; Schurig, V. J. Chromatgr. A 1996, 723, 399-404. (18) Pakdeesusuk, U. Ph.D. Dissertation, Clemson University, Clemson, SC, 2002. (19) Shelton, D. E.; Tiedje, J. M. Appl. Environ. Microbiol. 1984, 47, 850-857. (20) Pakdeesusuk, U.; Freedman, D. L.; Lee, C. M.; Coates, J. T. Environ. Toxicol. Chem. (in press). (21) Mazur, C. S.; Jones, W. J. Environ. Sci. Technol. 2001, 35, 47834788. (22) Quensen, J. F., III; Boyd, S. A.; Tiedje, J. M. Appl. Environ. Microbiol. 1990, 56, 2360-2369. (23) Dunnivant, F. M.; Elzerman, A. W. J. Assoc. Off. Anal. Chem. 1987, 71, 551-556. (24) Wong, C. S.; Garrison, A. W. J. Chromatogr. A. 2000, 866, 213-220. (25) Wong, C. S.; Hoekstra, P. F.; Karlsson, H.; Backus, S. M.; Mabury, S. A.; Muir, D. C. G. Chemosphere. 2002, 49, 1339-1347. (26) Kim, J.; Rhee, G.-Y. Environ. Toxicol. Chem. 1999, 18, 2696-2702. (27) Singer, A. C.; Wong, C. S.; Crowley, D. E. Appl. Environ. Microbiol. 2002, 68, 5756-5759. (28) Nardi-Dei, V.; Kurihara, T.; Park, C.; Esaki, N.; Soda, K. J. Bacteriol. 1997, 179, 4232-4238. (29) Nies, L.; Vogel, T. M. Appl. Environ. Microbiol. 1990, 56, 26122617. (30) Ye, D.; Quensen, J. F., III; Tiedje, J. M.; Boyd, S. A. Appl. Environ. Microbiol. 1995, 61, 2166-2171. (31) Williams, W. A. Environ. Sci. Technol. 1994, 28, 630-635. (32) Ye, D.; Quensen, J. F., III; Tiedje, J. M.; Boyd, S. A. Appl. Environ. Microbiol. 1999, 65, 327-329. (33) Ye, D.; Quensen, J. F., III; Tiedje, J. M.; Boyd, S. A. Appl. Environ. Microbiol. 1992, 58, 1110-1114. (34) Loffler, F. E.; Ritalahti, K. M.; Tiedje, J. M. Appl. Environ. Microbiol. 1997, 63, 4982-4985. (35) Zipper, C.; Fleischmann, T.; Kohler, H.-P. E. FEMS Microbiol. Ecol. 1999, 29, 197-204.

Received for review August 8, 2002. Revised manuscript received December 14, 2002. Accepted December 20, 2002. ES026039G VOL. 37, NO. 6, 2003 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

9

1107