Biodegradation of Chlorinated Ethenes at a Karst Site in Middle Tennessee

Biodegradation of Chlorinated Ethenes in the Shallow Water-Bearing Zone

Ground water in the shallow water-bearing zone flows along a trough in the bedrock surface, which serves as the main route for horizontal transport of chlorinated ethenes in the shallow zone. The shallow ground-water contamination plume moved laterally from well 33S to well 7S to well 2S to well 5S (fig. 11). Because the general direction of ground-water flow and the extent of the chlorinated-ethene plume in the shallow water-bearing zone were known, identifying discrete oxidation-reduction zones along the flow path in the shallow zone was possible. The delineation of oxidation-reduction zones, supplemented with organic chemistry data, was used to infer which chlorinated-ethene degradation processes prevailed along the plume.

Geochemical data for water samples collected from well 33S indicate that anaerobic conditions were present in shallow ground water beneath the manufacturing building (fig. 11). Compared to water samples from other shallow wells, samples collected from well 33S on August 22, 1997, contained low DO and NO₃^- concentrations (<0.1 and <0.06 mg/L, respectively) and high NH₃, Mn²⁺, and Fe²⁺ concentrations (1.6, 2.6, and 18 mg/L, respectively). Anaerobic conditions were present in water samples from well 33S (table 7), and concentrations of electron acceptors were within ranges normally associated with iron or possibly sulfate-reducing conditions. Those oxidation-reduction conditions are associated with reductive-dechlorination processes.

Chlorinated-ethene and degradation-product data from shallow wells under the building also support the inference that reductive dechlorination was occurring in shallow ground water beneath the manufacturing building (table 8). Water samples from well 33S contained lower molar concentrations of TCE than molar concentrations of cDCE and VC (reductive-dechlorination byproducts) and cDCE and VC concentrations generally increased in water samples collected between 1994 and 1997 (fig. 12). Ethene and ethane (degradation products of VC) also were detected in water samples from shallow wells screened beneath the manufacturing building. Water samples collected from well 33S contained 60 to 93 µM/L of ethene and 28 to 34 µM/L of ethane (table 7). Water samples from well 32S, also screened beneath the manufacturing building, contained 278 to 379 µM of ethene. A low molar ratio of parent material to byproducts and increasing concentrations of byproducts are evidence of biologically mediated reductive dechlorination of TCE.

As the shallow ground water moved downgradient from the building, direct recharge resulted in higher DO concentrations in the ground water. Water samples from shallow wells not screened beneath the manufacturing building commonly contained DO concentrations greater than 1.0 mg/L (fig. 11). Conditions in wells away from the manufacturing building, near well 5S for example, were consistently aerobic (table 7), which would preclude reductive dechlorination from occurring. Aerobic degradation (cometabolism) of chlorinated ethenes is possible if sufficient food and oxygen are available. Ethene and ethane, which were found in shallow ground water under the building, are suitable food substrates for inducing cometabolism. However, concentrations in water samples from wells to the west of the manufacturing building (5S, 6S, and 35S) were usually below detection limits.

A more consistent food supply would be required to sustain cometabolism in the aerobic zone. Ammonia is an alternative food substrate of the AMO cometabolic pathway. The anaerobic conditions underneath the manufacturing building resulted in elevated concentrations of NH₃ (approximately 0.5 to 2.5 mg/L) and low NO₃^- concentrations in water samples from wells near the manufacturing building (table 7). Farther downgradient, where conditions were consistently aerobic (near well 5S), NH₃ concentrations were lower (less than 0.10 mg/L) and NO₃^- concentrations increased (0.59 to 1.8 mg/L) indicating that NH₃ was oxidized to NO₃^- (fig. 13). Concurrent decreases in NH₃ concentrations and increases in NO₃^- concentrations along the delineated flow path are consistent with the AMO pathway and the cometabolic degradation of TCE, cDCE, and VC. These geochemical data indicate that cometabolism by the AMO pathway may have been occurring in shallow ground water to the west of the manufacturing building. Direct oxidation of VC and cDCE is another potential biodegradation pathway in the aerobic part of the shallow water-bearing zone, but this was not confirmed in this study.

Geochemical conditions of water samples from wells located near the western side of the manufacturing building (7S and 2S) changed in response to rain events. DO and NO₃^- concentrations decreased and NH₃ and Mn²⁺ concentrations increased significantly between November 12, 1996 and February 11, 1997 in water samples from well 7S (fig. 14), indicating a change from aerobic conditions to manganese-reducing conditions. The geochemical sampling technique used during the November 12, 1996 and February 11, 1997 sampling events may have resulted in some aeration of samples and actual DO concentrations were probably lower than the measured concentrations. The sampling technique was modified after the February 11, 1997 sampling event to minimize aeration of samples. The geochemical data indicate that at times the zone of anaerobic reductive dechlorination may have extended beyond the edge of the building; however, whether significant reductive dechlorination occurred in these fluctuating zones of transition is not clear.

Geochemical conditions in this fluctuating transition zone were probably affected by both infiltration of water from the surface after rainfall events (fig. 15) and transport of shallow ground water from the upgradient anaerobic zone beneath the manufacturing building. As the shallow ground water moves laterally out from under the building, recharge from rain would supply DO and dilute the contaminated water. Thus, the anaerobic water containing ammonia, small aliphatic hydrocarbons, and chlorinated ethenes coming from under the building mixes with aerobic waters, making the conditions suitable for cometabolism. Methanotrophic and ammonia-oxidizing bacteria would consume methane and ammonia as well as oxygen in this transition zone. Cometabolic destruction of chlorinated ethenes would also occur in this transition zone. Undoubtedly some of the water carrying bacteria, contaminants, electron donors, and electron acceptors in the shallow water-bearing zone also migrates down into the karst aquifer. Evidence that aerobic and anaerobic biodegradation pathways remained active in the karst aquifer is examined in the next section.

Biodegradation of Chlorinated Ethenes in the Karst Aquifer

Water-quality data collected during periodic sampling by the USGS indicate that conditions in water-producing zones of the upper part of the Ridley Limestone vary spatially and temporally at the study site. Conditions in some zones, such as those intersected by wells 1D and 3D, were consistently anaerobic, whereas conditions in other zones, such as those intersected by wells 2D and 12D, fluctuated between anaerobic and aerobic (table 9). When anaerobic conditions were present, Mn⁴⁺ or Fe³⁺ reduction may have been the dominant electron acceptor (fig. 16); however, sulfate reduction also may have been occurring. Sulfide concentrations ranging from 0.002 to 3.46 mg/L were detected in water samples from deep wells (table 9). The sulfate and sulfide concentrations were generally higher in samples from the karst aquifer compared to the shallow wells (tables 7 and 9). This is probably due in part to the availability of sulfate from the gypsum nodules of the karst bedrock. Bacteria can accelerate the remobilization of sulfate and convert the sulfate to sulfide.

Significant concentrations of chlorinated ethenes (TCE, cDCE, and VC) were detected in water samples from several deep wells (table 9). Those wells containing chlorinated ethenes, except for well 12D, had higher molar concentrations of reductive-dechlorination byproducts (cDCE and VC) than of TCE (fig. 17). The concentration of chlorinated ethenes varied with hydrologic conditions and weather events. Periodic increases in VC and cDCE concentrations in water samples were detected. For example, between May 10 and August 10, 1996, cDCE concentrations in water samples from wells 3D and 4D increased from 0.01 to 7.48 µM/L and from 0.04 to 5.41 µM/L, respectively (fig. 18). Ethene and ethane also were detected in water samples from wells 1D, 2D, and 3D (table 9). The ethene and ethane are final products of reductive dechlorination of VC and are readily consumed by bacteria.

The high ratio of degradation byproducts to TCE implies that reductive dechlorination is occurring at the site. The location of the reductive-dechlorination process is not clear from the chlorinated-ethene data. The degradation byproducts may have been formed in the shallow water-bearing zone and transported down into the upper part of the Ridley Limestone without any additional dechlorination occurring in the karst aquifer. Traditionally, sequential patterns of oxidation-reduction zones along a contaminant plume transect are used to identify where reductive dechlorination is active in the aquifer. A simple contaminant flow path and sequential oxidation-reduction zones could not be identified for the karst aquifer because of the complex hydrogeology at the study site. The geochemical and biological information derived from sampling karst-aquifer wells were considered individually instead of as a section of a continuum along a flow path. The data for an individual well represents geochemical and biological conditions in that part of the karst aquifer, whether the aquifer is in an active or less active flow zone. Multiple lines of biodegradation evidence are presented for karst wells 12D, 1D, 3D, and 2D in the following text.

Multiple Lines of Evidence from Well 12D

Water-quality conditions of samples from well 12D, located near Snell Branch and screened in the upper water-producing zone of the Ridley Limestone, changed significantly after rainfall events. Continuous data collected from well 12D indicated that water levels rose (sometimes to land surface), specific conductance decreased due to dilution by rainwater, and DO concentrations and ORP (expressed as Eh) increased (fig. 19). These changes typically occurred within hours of rainfall events (fig. 20) indicating that water was quickly transported from land surface to the upper water-producing zone of the Ridley Limestone intersected by well 12D. These changes in water chemistry are consistent with the detection of a mud-filled cavity that was hydraulically connected to Snell Branch during the construction of well 12D. After rainfall events, the DO would start to decrease and anaerobic conditions would develop until the next rainfall (fig. 19).

During dry periods such as late May 1998, anaerobic conditions persisted and oxidation-reduction potentials decreased (fig. 19D) to levels associated with iron-reducing conditions (fig. 2). However, during the 3 months of continuous monitoring, anaerobic conditions did not persist long enough for reductive dechlorination to occur. The lack of VC (fig. 21), ethene, and ethane (table 9) in water samples from well 12D also suggests that little reductive dechlorination had occurred. Some periodic increases in cDCE concentrations were detected in water samples collected from well 12D; however, these increases were often accompanied by increases in TCE (fig. 21) and are most likely the result of transport processes, not reductive dechlorination in the water-producing zone intersected by well 12D.

These results indicate that natural attenuation of chlorinated ethenes in the part of the karst aquifer intersected by well 12D is probably limited to aerobic degradation or mechanisms such as dilution. Ammonia-oxidizing bacteria and methanotrophs (table 10) were detected in water samples from well 12D; however, microcosms containing water from well 12D (experiment 2, treatment 4) did not exhibit significant aerobic degradation of TCE during a 17-week period (fig. 22). Microcosm data are given in table 11. The water samples used to construct these microcosms were collected during a dry period (May 21, 1998) and contained unusually low DO concentrations (table 9) and ORP (fig. 19D), which limited aerobic degradation. Between weeks 17 and 23, some reduction in TCE concentrations occurred in excess of the reductions in control samples (fig. 22) which indicate that reductive dechlorination was beginning to occur.

Data from microcosm experiment 1 indicate the occurrence of reductive dechlorination in microcosms containing water from well 12D. During microcosm experiment 1, the discovery was made that control microcosms (experiment 1, treatment 9) were not sterile. Plate counts for facultative and aerobic heterotrophic bacteria indicated greater than 2 x 10⁴ bacteria colony-forming units per 100 milliliters (CFU/100mL) of water sample in the control microcosms after a 1-week incubation period. Bacteria enumeration data are given in table 12.

Without adequate controls, determining to what degree the TCE decreases in the microcosms was caused by biodegradation was impossible. Significant cDCE and VC concentrations, however, were detected in several of the treatments after a 10-month incubation period, indicating that some reductive dechlorination had occurred in the microcosms. VC and cDCE were not present in the microcosms when they were first constructed, and the detection of the compounds could only be due to reductive dechlorination of TCE.

After a 10-month incubation period, cDCE represented 98 and 67 percent of the total chlorinated-ethene concentration remaining in microcosm treatments containing water from well 12D (experiment 1, replicates 7 and 8, respectively) (fig. 23). Microorganisms responsible for reductive dechlorination (sulfate reducers) were detected in water samples from well 12D (table 10), and the microcosms from experiment 1 indicate that the microorganisms were able to dechlorinate TCE after oxygen and other electron acceptors were depleted during the 10-month incubation period. In spite of microcosm evidence that reductive dechlorination can occur in water from 12D, it is unlikely that reductive dechlorination is a significant biological process in that part of the karst aquifer. Field geochemical conditions indicate that the reducing conditions in 12D are disrupted with each rainfall event.

Multiple Lines of Evidence from Well 1D

Well 1D is screened in the lower water-producing zone of the karst aquifer and is located close to the manufacturing building. Water-quality conditions in the part of the karst aquifer intersected by well 1D exhibited little change in response to storms during the 2 months (March 1998 to May 1998) of continuous monitoring. Anaerobic conditions (less than 0.3 mg/L DO) were consistently detected in water samples collected from well 1D during quarterly sampling events (table 9), and geochemical data indicate manganese-reducing conditions (fig. 16). Continuous water-quality monitoring indicated that although water levels increased after precipitation events, specific conductance, DO, and ORP (expressed as Eh) changed little (fig. 24). Equilibrium of the ORP electrode took approximately 4 days to come to equilibrium with the aquifer water. Once equilibrium was reached, ORP measured during continuous monitoring was normally between -150 and -200 millivolts (mV), which is normally associated with nitrate-, iron-, and sulfate-reducing conditions (figs. 2 and 24D).

During a 23-week period, significant aerobic degradation of TCE was not detected in microcosms using water collected from well 1D on May 1, 1998 (experiment 2, treatment 1) (fig. 22). Ammonia oxidizers were not detected in water samples collected from well 1D; however, methanotrophs were detected (table 10). Based on the geochemical data and the microcosm results, biodegradation of chlorinated ethenes in the water-producing zone intersected by well 1D would be limited to reductive dechlorination. Between 17 and 23 weeks, TCE concentrations in the microcosms using well 1D water decreased faster than in the control samples (fig. 22) which could indicate that reductive dechlorination was beginning to occur. During the 10-month incubation period of microcosm experiment 1, TCE degradation byproducts (VC and cDCE) increased from zero to approximately 80 and 50 percent of the average total chlorinated-ethene concentration remaining in microcosms containing water from well 1D (experiment 1, treatments 1 and 2, respectively) (fig. 23). Results from microcosm experiment 1 demonstrate that reductive dechlorination of TCE is possible in the water from well 1D.

The lag time for reductive dechlorination to occur in the microcosms was probably due to aeration of the experimental water during set up. No lag time would be expected in the aquifer around well 1D since the conditions are consistently anaerobic. Further evidence of reductive dechlorination in well 1D is found in the chlorinated-ethene data (table 8). Byproducts of TCE reductive dechlorination, DCE and VC, were found in every sample reported. Thus, the multiple lines of evidence indicate reductive dechlorination occurs in the part of the karst aquifer intersected by well 1D.

Multiple Lines of Evidence from Well 3D

Well 3D is located near the center of the manufacturing building and is screened in the lower water-producing zone of the upper part of the Ridley Limestone. Continuous water-quality data indicate that rainfall events and changes in the pumping rate of well 9D affect water-quality conditions in the part of the karst aquifer intersected by 3D. When pump-and-treat well 9D was operating (March 19, 1998 to April 8, 1998) the water moving through the screened interval of well 3D appeared to have been stored in the karst aquifer longer than the water moving through the screened interval of well 12D. Evidence for this inference includes increased specific conductance after precipitation events (fig. 25B). The higher specific conductance is indicative of increased dissolved solids normally associated with limestone dissolution over time. Also, dye was detected in water samples collected from well 3D in 1996, 5 years after the dye was injected into well 8D which was screened in the same lower water-producing zone of the karst aquifer. Water samples collected from well 3D during quarterly sampling normally had low DO concentrations (less than 0.2 mg/L) (table 9). During the continuous monitoring period, DO concentrations in the well water were less than 0.5 mg/L and ORP was less than -100 mV (fig. 25D) when the pump-and-treat well was operating.

The ORP (fig. 25D) was within ranges normally associated with nitrate-, iron-, and sulfate-reducing conditions (Stumm and Morgan, 1981). Sulfide concentrations detected in some water samples from well 3D (>1 mg/L) were significantly higher than sulfide concentrations detected in water samples from shallow wells (tables 7 and 9), which indicate sulfate-reduction occurred in the aquifer near well 3D. Sulfate-reducing bacteria were identified in water samples collected from well 3D (table 10). Results from microcosm experiment 1 indicate that bacteria from well 3D can reductively dechlorinate TCE. After a 10-month incubation period, cDCE represented 98 and 24 percent of the total chlorinated ethene remaining in microcosms using well 3D water (experiment 1, treatments 5 and 6, respectively) (fig. 23).

The specific conductance of water in well 3D decreased, and DO increased, when pump-and-treat well 9D was not operating (various times between April 8, 1998 and May 20, 1998) (fig. 25). The slight decrease in specific conductance and increased DO associated with pump-and-treat downtime indicate recharge of the karst aquifer near well 3D by fresh rainwater. The resulting change would tend to benefit the aerobic bacteria, ammonia oxidizers and methanotrophs, detected in water samples from well 3D. Results from the second microcosm experiment using water collected on May 21, 1998, suggest that significant cometabolism occurred when aerobic conditions prevailed. During a 3-week incubation period, TCE in the microcosms containing well 3D water (experiment 2, treatment 3) was completely degraded (fig. 22).

The multiple lines of evidence combined indicate that aerobic and anaerobic degradation occurred in the water-producing zone of the karst aquifer intersected by well 3D. Anaerobic conditions were present with the pump-and-treat well operating (the normal condition during this study) and reductive dechlorination was likely. When pump-and-treat well 9D was turned off, aerobic conditions suitable for aerobic cometabolic degradation of chlorinated ethenes prevailed. Bacteria suitable for both anaerobic and aerobic degradation were present in well 3D, and evidence of both degradation pathways was found in the microcosm studies.

Multiple Lines of Evidence from Well 2D

Well 2D is located to the west of the manufacturing building near pump-and-treat well 9D and is screened in the lower water-producing zone of the karst aquifer, as is well 9D. Because well 2D is located very close to pump-and-treat well 9D, water passing through the screened interval of well 2D probably flows from several parts of the karst aquifer in response to the gradient imposed by pumping well 9D. Water in well 2D represents a composite of waters from parts of the karst aquifer affected by the pumping. The ORP and DO conditions in well 2D tended to be moderate whereas the bacteria composition ranged from aerobic to anaerobic bacteria types.

Water-quality conditions in the part of the karst aquifer intersected by well 2D changed in response to rainfall events, although not as quickly as conditions in well 12D. Specific conductance decreased after rainfall events due to dilution by rainwater (fig. 26B). DO concentrations initially increased after rainfall events then rapidly decreased as anaerobic aquifer water with a high chemical oxygen demand mixed with the rainwater (fig. 26C). In a pattern similar to the DO concentrations, ORP decreased after rainfall events (fig. 26D). These data indicate that freshwater was initially transported toward the pump-and-treat well after rainfall events (fig. 26). However, this fresh rainwater was soon followed by a mixture of anaerobic aquifer water and rainwater as indicated by the lower specific conductance, DO, and ORP concentrations. Presumably, old aquifer water displaced by the rainwater contained chemicals such as sulfide and reduced-metals (Fe²⁺ and Mn²⁺) that would scavenge DO carried in by the rainwater, also driving the ORP down. When the rain ceased, the pre-rain event flow pattern would return to the karst aquifer within days. The implication is that the hydrology in the karst aquifer was stable during dry periods punctuated by recharge events that carried in water, DO, food, and other compounds. Bacteria as diverse as those found in the karst aquifer would take advantage of these changing environments, switching between anaerobic and aerobic conditions.

The quarterly geochemical data (table 9) and continuous ORP monitoring (approximately +200 mV) (fig. 27D) indicate that baseline conditions ranged from slightly aerobic to Mn⁴⁺ and NO₃- reducing. Aerobic bacteria and NO₃^--reducing bacteria were identified in these water samples. During the anaerobic periods after a rain event, ORP commonly decreased to sulfate-reducing conditions (fig. 27D), and SO₄-reducing bacteria were detected in water samples from well 2D (table 10). However, the sulfate-reducing conditions lasted only a few days in well 2D, which would be insufficient time for significant reductive dechlorination to take place. The ORP conditions and bacteria types found in well 2D represent the wide range of conditions present in the karst aquifer. The pattern of ORP and DO with regard to baseline and post-rain conditions provides insight into the displacement of older ground water by recharge from rainwater.

Cometabolism appeared to be a more significant biological degradation process than reductive dechlorination in microcosms set up with water from well 2D. Experiment 2 microcosms using water from well 2D exhibited significant aerobic degradation. Within 3 weeks, TCE in the microcosms was completely degraded (fig. 22, treatment 2). Methanotrophs were identified in well 2D water samples; however, ammonia oxidizers were not detected. Bacteria identification and microcosm results in conjunction with geochemical monitoring indicate that environmental conditions suitable for cometabolism occurred more consistently in well 2D than did conditions for reductive dechlorination. However, results from microcosms in experiment 1 indicate that native bacteria from well 2D were capable of reductively dechlorinating TCE. Evidence of reductive dechlorination was found in microcosms after a 10-month incubation period. The TCE degradation byproduct cDCE was present in microcosms as a result of reductive dechlorination (experiment 1, treatments 3 and 4) (fig. 23).

Data Collected from Other Deep Wells

Geochemical data from other deep wells at the site indicated a similar pattern of anaerobic zones in the aquifer or zones fluctuating between aerobic and anaerobic conditions. The DO concentration in water samples from well 4D ranged from 0.2 to 1.2 mg/L and the DO concentration in water samples from well 16D-B ranged from 0.3 to 1.8 mg/L (table 9). Both aerobic (ammonia oxidizers and methanotrophs) and anaerobic (sulfate reducers) bacteria were detected in water samples from well 16D-B (table 10).

Water samples from monitoring wells 10D-B and 11D contained less than 0.1 mg/L of DO in three or fewer sampling events. Well 11D is screened in a bedding plane that appeared to be isolated from the active ground-water-flow system. Water from well 11D had such high sulfide concentrations that it had a strong odor and would start to form a black precipitate within minutes after being brought to the surface. These observations suggest that DO concentrations in well 11D were consistently less than 0.1 mg/L. Unlike well 11D, well 10D-B was screened in a highly fractured area of the karst aquifer. Limited geochemical data (table 9) indicate anaerobic conditions occurred in well 10D-B; however, both aerobic and anaerobic bacteria were present in water samples (table 10), implying a close association with aerobic conditions. More data are required to determine whether conditions in 10D-B are consistently anaerobic.

Geochemical data from 11 deep wells indicate that geochemical conditions varied greatly throughout the karst aquifer at the site (table 9). For example, during the August 20-22, 1997 sampling event, DO ranged from less than 0.1 to 5.5 mg/L, ammonia ranged from 0.02 to 2.30 mg/L, and sulfide ranged from 0.02 to 3.46 mg/L.

A large diverse microbial population was also present throughout the karst aquifer (table 10). The concentration of viable bacteria present in water samples from the deep wells was higher than expected. For example, aerobic heterotrophic bacteria concentrations in water samples from the karst aquifer were greater than 6 x 10⁴ colony forming units per 100 mL of sample (table 12). This range of bacteria does not account for strict anaerobic bacteria or others that would not grow on the tryptic soy agar plates. Still, these un-enriched bacteria concentrations occurring in the karst aquifer were equivalent to culture concentrations used in laboratory biodegradation studies (Dalton, 1977; Nelson and others, 1988). This suggests enough bacteria are present in the karst aquifer to support TCE biodegradation.

The hydrologic, biological, and chemical data demonstrate the large range of hydrologic conditions in the karst aquifer and provide insight into its complexity. Some areas in the karst aquifer responded rapidly to rainfall events whereas other areas were more stable and changed little regardless of rainfall, probably reflecting longer ground-water residence times. This range of hydrologic conditions provides an opportunity for bacteria to enter, reproduce, and spread throughout the aquifer system. The bacteria appear to have flourished wherever conditions were right. As the bacteria multiplied, they influenced the geochemistry of the system and degraded chlorinated solvents. Despite the complexity of the karst system, the same biological degradation processes are active in this karst system as in any unconsolidated aquifer. The hydrologic, biological, and chemical data provide multiple lines of evidence, supported by laboratory microcosm results, that a variety of TCE biodegradation pathways are active in the karst aquifer (table 13). The central issue for bioremediation is not whether biodegradation is occurring in the karst aquifer, but whether water is retained long enough in the karst aquifer to degrade a meaningful mass of contaminants. The observation of dyes in various deep wells 5 years after injection and the stable anaerobic conditions in parts of the aquifer indicate long local residence times. This study, however, does not determine what volume of the aquifer is represented by such areas and how much contamination they contain.