Site Description

Site K is a 21-acre area that is part of OU2 at TCAAP, the former New Brighton/Arden Hills Superfund site (fig. 1). The contaminant plume (fig. 2) is in a shallow perched aquifer, containing TCE and DCE (primarily cisDCE), as the contaminants of concern in the plume. The perched aquifer contains alluvial sand of the Fridley Formation, called Unit 1 at TCAAP, that has been disturbed in some areas of Site K by past remediation activities (refer to the “Previous Remedial Investigations and Activities” section). Treated water from the air stripper plant, which has been operating since 1986, is discharged into Rice Creek. Water-quality and water-level monitoring downgradient from the collection trench indicates that the approximately 18-foot (ft) deep trench typically captures the groundwater plume in Unit 1, which is upgradient from the trench and beneath the former Building 103 footprint (fig. 2; PIKA Arcadis, 2019). Four wells near the trench (01U625D, 01U626D, 01U627D, and 01U628D) that are screened at the interface of Unit 1 and the underlying aquitard, Unit 2 (fig. 2), were sampled in March 2000 to determine the potential for DNAPL migration beneath the trench along the Unit 1 aquifer base, and no DNAPL presence was found (PIKA Arcadis, 2019). A sentinel well (03U621) was incorporated by the Army in the TCAAP monitoring plan to monitor potential VOC migration through the Unit 2 aquitard and into the underlying aquifer at Site K. No VOCs were detected in the sentinel well; 1,4-dioxane has been detected but is not considered to result from remedial activities at Site K (PIKA Arcadis, 2019). Under the OU2 1997 ROD, the extraction trench and air stripper treatment at Site K will continue to be implemented until groundwater cleanup levels of 30 μg/L for TCE and 70 μg/L for total DCE are attained. Treated water discharges to the storm sewer that outlets to Rice Creek.

Previous Remedial Investigations and Activities

The history of remedial activities at Site K is summarized in figure 3. After construction of Building 103 was completed in 1942, it was used to produce small caliber ammunition, operating until 1945 and then again in 1950–51 (Conestoga-Rovers & Associates, Inc., 2009² ). Between 1946 and 1950, Building 103 was used to clean and store production machinery. The building was inactive after 1951 until reactivated from 1961 to 1998 for the production of fuses, mines, and weapons systems by Honeywell and its successor, Alliant Techsystems, Inc. Materials used in former Building 103 included chlorinated solvents and degreasers (TCE, 1,1,1-trichloroethane, and dichlorofluoromethane), radioactive materials, and petroleum products (xylene, toluene; Conestoga-Rovers & Associates, Inc., 2009). Contamination from Building 103 was first detected in 1983 and consisted of TCE, oil, and cyanide in stormwater that was discharging into Rice Creek (Conestoga-Rovers & Associates Limited, 1984²). Subsequent investigations identified the groundwater VOC plume in Unit 1 at Site K, but a soil source was not found (Conestoga-Rovers & Associates, Inc., 2009). Thus, the OU2 1997 ROD stipulated additional characterization of the unsaturated soil of Unit 1 at Site K to locate the source area of the groundwater VOC contamination.

Soil and soil gas sampling for VOCs was conducted in 2000 with direct-push probes, leading to the identification of a sump on the east side of Building 103 (fig. 4) as the source of the Site K groundwater contamination (Conestoga-Rovers & Associates, Inc., 2009). In 2008, remediation goals for contaminated soil were set at 46 milligrams per kilogram (mg/kg) for TCE and 22 mg/kg for cisDCE, which were the industrial Soil Reference Values established by the Minnesota Pollution Control Agency for the two identified contaminants of concern (Conestoga-Rovers & Associates, Inc., 2009); soil excavation and off-site disposal was the recommended remedial alternative (Conestoga-Rovers & Associates, Inc., 2009). After the signing of an Action Memorandum (October 16, 2008), the recommended soil remedy, which included the removal of 41 tons of soil and additional rubble and clean concrete in a 20-square-foot (ft²) area of the sump and an additional 112-ft² area to the east, was completed in June–August 2009 (Conestoga-Rovers & Associates, Inc., 2009).

In the vicinity of the former sump, soil excavation was initially done to the water table at a depth of about 6 feet below ground surface (ft bgs) but was then extended to the clay at the bottom of Unit 1 at a depth of about 8 ft bgs. Although not required as part of the removal action or ROD for Site K, 450 lb of granular potassium permanganate (KMnO₄) were added at the bottom of the 2009 excavation area before covering, in an attempt to enhance contaminant mass removal of VOCs in the saturated zone by chemical oxidation (Conestoga-Rovers & Associates, Inc., 2009; Wenck Associates, Inc. and others, 2010, 2011²). Excavated site soil that was determined to be clean was used in backfilling. Post-excavation testing determined that the unsaturated soils in the excavation area were acceptable for unrestricted use (Conestoga-Rovers & Associates, Inc., 2009), and a ROD amendment stated that no further action regarding the soil was necessary (U.S. Army and others, 2012¹).

Groundwater samples were collected from the bottom of the excavation and from two downgradient wells (01U611 and 01U609) before and after the addition of the KMnO₄. Groundwater TCE concentrations that had pooled in the excavation pit decreased from 710 μg/L to below detection within 2 hours after KMnO₄ addition. However, the concentrations in wells 01U611 and 01U609 did not show a decreasing trend between baseline conditions (June 23, 2009) and samples collected 4 months (October 2009) and 12 months (June 2010) post-KMnO₄ treatment (Conestoga-Rovers & Associates, Inc., 2009; Wenck Associates, Inc. and others, 2011).

Additional activities at Site K have included removal of Building 103’s foundation slabs and utility infrastructure in 2013–14 and targeted soil excavation of suspected sources in 2014 under a Response Action Plan (Wenck Associates, Inc., 2016). These activities were referred to as “site redevelopment activities” and were completed by Ramsey County after acquiring a 427-acre portion of TCAAP, including Site K, on April 15, 2013. Before the slab at the northern end of Building 103 was removed, it was temporarily used to stage and compost petroleum-contaminated soil excavated from various locations on the TCAAP property (Wenck Associates, Inc., 2016). Soil excavation was done in November 2014 and removed approximately 7,000 cubic yards of VOC-effected soil from the former Building 103 area, including soil in and downgradient from the former sump area. At the time of excavation, the water table was about 5 ft bgs, and soil was excavated to a depth of about 0.5–1 ft below the water table. Figure 4 shows the 2014 excavation footprint and locations where soil VOC contamination was known to be deeper than the 6 ft excavation depth, based on 2007 soil sampling by TetraTech and earlier sampling (Wenck Associates, Inc., 2016). This contaminated soil was left in place. Additional excavation was completed along stormwater piping, in a linear area west of the sump removal area (fig. 4).

Removal of the stormwater piping may have allowed an increase in recharge and a rise in the water table in the aquifer. Prior to 2014, the contaminated aquifer in the former building area was only 1- to 2-ft thick; it increased to about 5- to 6-ft thick after 2014 (PIKA Arcadis, 2019).

Historical Plume Concentrations and Movement

In the plume in Unit 1 at Site K, historical TCE concentrations were generally highest in wells 01U609 and 01U611, which were located within the former Building 103 footprint and near the former sump area (fig. 4), although 01U609 was not part of the monitoring plan for yearly sampling. In the fiscal year 2018 Annual Performance Report, TCE concentrations in the Site K plume were a maximum of 1,500 micrograms per liter (µg/L), but several wells monitored in the core of the plume, including 01U611, were removed in 2014 (PIKA Arcadis, 2019). Prior to 2014, 01U611 had a maximum concentration of 11,000 μg/L (PIKA Arcadis, 2019). Historical maps of TCE concentrations placed both 01U608 and 01U609 inside the 1,000 µg/L contour for the plume (fig. 4), and 01U609 had particularly high concentrations. In 2000, TCE concentrations in samples from well 01U609 ranged from 61,000 to 110,000 µg/L (Sola, 2001¹). Thus, both 01U609 and 01U611 had historical concentrations above 1 percent of TCE solubility in water (or 1,100 milligrams per liter [mg/L] at 20 degrees Celsius [°C]), which indicates potential presence of groundwater contact with nearby DNAPL (Kueper and Davies, 2009). Abandoned wells 01U608, 01U609, and 01U611 were replaced by the USGS in 2020–21 as part of the treatability study. These wells are listed as 01U608R, 01U609R, and 01U611R in table 1.

In addition to TCE, historical data show that the plume at Site K contained cisDCE and trans-1,2-dichloroethene (transDCE), which are likely produced by reductive dechlorination of TCE. However, cisDCE and transDCE were generally at concentrations more than an order of magnitude lower than the TCE concentrations, and VC generally was undetected. There are few historical measurements of constituents that would indicate the redox conditions in the plume.

The Site K plume trends northwest toward the excavation trench (fig. 4). Water-quality and water-level monitoring downgradient from the trench indicates that the trench typically captures the plume groundwater. An exception is that the TCE concentration in the downgradient well 01U603 was 5,600 μg/L in 2014, whereas TCE was not previously detected in this well (PIKA Arcadis, 2019). TCE concentrations decreased to 5.1 μg/L in fiscal year 2018 (PIKA Arcadis, 2019). High groundwater levels in April and May 2014 are believed to have mobilized TCE DNAPL in the former storm sewer bedding that was present underneath the former building footprint (PIKA Arcadis, 2019). Removal of the stormwater piping in 2014 may have allowed an increase in recharge and a rise in the water table elevation of the aquifer, as reported through 2018 (PIKA Arcadis, 2019). Historical water levels in 01U608, 01U609, and 01U611 were a maximum of 886.83–888.06 ft above mean sea level (PIKA Arcadis, 2019).

Previous Bioremediation Studies

Two pilot studies of enhanced bioremediation, both designed to promote reducing conditions and a subsequent increase in reductive dechlorination of TCE, were performed in the perched aquifer at Site K in 2000 (Sola, 2001). The locations of the studies’ test injection wells are shown in figure 4. The objective of the studies was to determine if bioremediation could be applied to the plume, leaving only the source area soils to be remediated and allowing the groundwater extraction system to be shut down. The field tests were conducted prior to the demolition of Building 103 (2006) and removal of the concrete and stormwater piping (2012–15).

The first test consisted of injection of 2,700 lb of commercially available Hydrogen Release Compound (HRC) in a 65-by-70-ft area of the aquifer beneath Building 103 (fig. 4), distributed through 89 injection points using a high-pressure pump. Although drilling showed that the aquifer was thinner than expected (about 2 ft thick), the injection location was not moved closer to the excavation trench, where the aquifer was known to be over 15-ft thick, because the goal was to treat the plume underlying the building. The results from monitoring conducted May–November 2000 showed that TCE concentrations in the test area did not change after HRC injection. Four monitoring wells were installed for performance sampling within and downgradient from the injection area, but one of these wells had insufficient water for sampling. Well 01U609 was upgradient from the test area and was used as a background monitoring location, although VOC concentrations were much higher in this well (61,000–110,000 μg/L) than in the injection area (140–1,600 μg/L) during the test period (Sola, 2001). The unexpectedly thin aquifer beneath the building area was believed to have caused the failure of the HRC injection test, either because the HRC was insufficiently hydrated or lost into the vadose zone (Sola, 2001).

The second test was conducted outside of Building 103, in the downgradient edge of the plume between well 01U615 and the trench, where the thickness of the perched aquifer extended to a depth of 24 ft bgs (Sola, 2001). This test consisted of hydrogen delivery to the aquifer through experimental gas permeable membranes, using an array of 16 injection and monitoring points in 3 rows within an 8-by-4-ft area. The goal of this test was to evaluate the use of a permeable bioremediation barrier to treat the groundwater transported from the source area as a potentially more cost-effective plume treatment than the groundwater extraction system.

The hydrogen was delivered in the first upgradient row of the array (fig. 4). Two additional monitoring wells were installed and sampled downgradient from the array, and well 01U615, located 8-ft upgradient from the first row of the injection array, was the planned background monitoring location. Natural advection and diffusion were found to distribute the hydrogen throughout the array and upgradient from well 01U615. The hydrogen gas test successfully showed a decrease in TCE concentrations and an increase in ethene, which indicated that complete reductive dechlorination of TCE was achieved. Thus, the hydrogen test indicated that native bacteria in the aquifer downgradient from the source area are capable of complete biodegradation of TCE, if stimulated using an electron donor. The bioremediation method, however, was determined to be relatively costly (Sola, 2001).

In addition to biotic reductive dechlorination, a natural attenuation microcosm study was conducted for the adjacent Building 102 plume and showed that abiotic degradation of cVOCs occurred at substantial rates (PIKA Arcadis, 2019). This abiotic degradation was believed to be associated with a reduction of iron and manganese oxides present in the sands in the aquifer, including magnetite. Abiotic degradation associated with magnetite was shown to be significant for cisDCE and 1,1-dichloroethene (11DCE) in the Building 102 plume (Ferrey and others, 2004). Similar abiotic degradation rates were determined for cisDCE and 11DCE in plumes in the glacial outwash sand aquifer at Site A at TCAAP (Ferrey and others, 2004). These sites had substantially lower TCE and DCE concentrations than the Site K plume, however.

Pre-Design Site Characterization

Soil removal and site redevelopment activities at Site K through 2014 led to the abandonment of wells and a lack of soil and groundwater data in the core of the plume and near the suspected source area for the TCE after those activities. Soil coring, well installation, groundwater sample collection, and hydrologic measurements were completed during December 2020–July 2021 by the USGS to assist in developing a current conceptual site model (CSM) and to collect data and site materials needed for designing the field bioremediation test.

A direct-push coring rig was used to drill replacements of previously abandoned wells 01U608, 01U609, and 01U611 in the source area (table 1). Soil cores were collected in acetate liners at each location, recovering 1.75-inch (in.)-diameter cores to the base of the perched aquifer (Unit 1) to describe lithology and determine well screen placement. Intact soil cores were then collected in an adjacent hole at each location and used for well installation. Intact cores were capped, placed on ice, and sent overnight to the USGS Maryland-Delaware-DC Water Science Center (MD-DE-DC WSC) for sectioning of cores to obtain soil samples for VOC analyses. Soil samples (approximately 5 grams each) were placed in pre-weighed VOC vials containing 10 milliliters (mL) of methanol and stored in the refrigerator for a minimum of 48 hours of contact time. The methanol extract was then transferred to VOC vials and analyzed for VOCs within 14 days. The method used was the same as that applied to the water samples and is described in the “Groundwater Sampling and Geochemical Analyses” section. Remaining soil from the cores was transferred to mason jars that were filled to leave minimum headspace and refrigerated for use in microcosm experiments. Sandy soil from the aquifer was kept separately from clayey layers near the base of the aquifer.

Stainless-steel piezometers with 2-ft-long stainless-steel wire-wrapped screens were installed in the replacement wells 01U608R, 01U609R, and 01U611R (table 1). The piezometers were left undisturbed for a minimum of 24 hours after installation before development occurred. The piezometers were developed by removing a minimum of three casing volumes of water from the well using a peristaltic pump. Water-quality parameters, including temperature, pH, dissolved oxygen, and specific conductance, were measured and recorded using a water-quality meter during well development. Measurements were taken periodically during the purging process until the parameters stabilized. Sampling occurred 24 hours or more after well development.

Groundwater samples were collected from the replacement wells and from two existing wells (01U603 and 01U615) for the initial site characterization. Groundwater was collected using a peristaltic pump and dedicated tubing. Samples were analyzed for VOCs, field parameters, redox-sensitive constituents (nitrate, ammonia, sulfide, ferrous iron [Fe²⁺], methane), dissolved organic carbon, major ions, and metals using methods summarized in the “Groundwater Sampling and Geochemical Analyses” section. Groundwater was also collected from 01U611R, 01U603, and 01U615 for use in microcosm experiments described in the “Microcosm Experiments” section. Unfiltered water was collected for experiments in glass mason jars that were sterilized by autoclaving, rinsed with groundwater before filling with minimum headspace, and then refrigerated until use. Initial site characterization also included water-level measurements and slug tests of the newly replaced wells to obtain data for evaluation of groundwater flow directions and transmissivity in the perched aquifer. Hydrologic methods used in this initial site characterization are the same as those described in the “Performance Monitoring” section.

Microcosm Experiments

A series of anaerobic microcosm experiments was conducted to determine the natural degradation potential in the aquifer and to determine the amendments needed to enhance TCE biodegradation at the site (table 2). All treatments were prepared in an anaerobic chamber in 160-mL glass serum bottles closed with crimped butyl stoppers and contained 100 milliliters (mL) of liquid and 10 mL of solids. Water used in the microcosm treatments was either site groundwater or simulated groundwater, depending on the objective of the experiment. Simulated groundwater had a pH of 7.0 and contained 8.5 mg/L KH₂PO₄, 22 mg/L K₂HPO₄, 33 mg/L Na₂HPO₄, 10 mg/L Ca₃(PO₄)₂, 1.25 mg/L FeCl₃·6H₂O, 18.5 mg/L (NH₄)₂SO₄, and 25 mg/L MgSO₄·7H₂O. Solids used in the experiments included autoclaved sand for controls, sand incubated in site wells in passive microbial samplers (PMSs), or soil from cores collected at the site. Soil core samples were passed through a 2-millimeter sieve in an anaerobic glove chamber and thoroughly mixed before use in experiments. Microcosm experiments included treatments to evaluate different fast- and slow-release electron donors to support reductive dechlorination of TCE and amendments containing the WBC-2 dehalogenating culture to evaluate bioaugmentation. WBC-2 was obtained from SiREM Lab (Guelph, Ontario, Canada), where it is grown and maintained in large volumes.

TCE was amended to all treatments to a nominal aqueous-phase concentration of 5 mg/L at the start of experiments (day 0) to achieve comparable aqueous concentrations in all treatments. Site water was purged using nitrogen gas to remove VOCs before microcosm preparation; however, VOCs were also present in the site soil used in many treatments.

For the first experiment (PMS experiment), the potential for anaerobic TCE biodegradation by the native microbial community in the perched aquifer was evaluated. PMSs were deployed for 8 weeks in monitoring wells 01U611R, 01U603, and 01U615 to obtain the native microbial community that prefers attachment to surfaces (fig. 4; table 1). The PMSs, which were tested as part of a previous study (Allen-King and others, 2021), were constructed in the USGS MD-DE-DC WSC of a fine stainless-steel mesh filled with clean drillers sand that was sieved to remove fines and autoclaved. The mesh was cinched using stainless-steel wire, and the assembled microbial samplers were autoclaved before wrapping in aluminum foil for storage and transport to the site. Two microbial samplers were linked sequentially and suspended for 8 weeks within and directly above the screened interval in each well using nylon string, allowing colonization of the sand by in situ microbial populations before retrieval. Upon retrieval, intact PMSs were placed immediately in autoclaved wide-mouth glass mason jars that were filled with groundwater from the respective well, which would contain the suspended microbial community and provide representative site geochemistry for the experiments. Jars were filled until minimal headspace remained and shipped on ice overnight to the USGS MD-DE-DC WSC laboratory. Microcosm treatment sets were constructed with the sand from the incubated PMSs and the associated water for each well and included treatments bioaugmented with WBC-2 (10 percent by volume). A control was prepared with autoclaved sand and deionized water (table 2). All microcosms were amended with 5 millimoles per liter (mmol/L) lactate, added as sodium lactate, as an electron donor.

In the second experiment, different electron donors (also known as carbon donors or substrates) were compared (table 2). Fermentation of the carbon donors releases hydrogen, which is the electron donor ultimately utilized by dechlorinating bacteria. Carbon donors were amended to achieve comparable hydrogen production ratios (Henry, 2010). The electron donors are designated as either fast-releasing compounds, which require continuous application or multiple applications at a site, or as slow-releasing compounds, which require fewer injections. This study tested lactate and corn syrup as fast-releasing donors and whey, soy-based vegetable oil (VO), and 3-D Microemulsion³ (3D; Regenesis, San Clemente, California) as slow-releasing donors. Lactate is the primary donor used to maintain and grow WBC-2 and, thus, provides a good baseline for comparison of WBC-2 activity for the other donors. Microcosms were constructed with soil collected from the perched aquifer at site 01U611R and simulated groundwater. Simulated groundwater was used for this experiment to eliminate possible matrix interference effects while comparing the donors. Each electron donor was tested in microcosms with and without added WBC-2 (10 percent by volume) to evaluate the need for bioaugmentation.

An additional microcosm experiment was conducted to evaluate if altering the timing of the addition of the electron donors and WBC-2 would increase degradation rates and decrease the accumulation of cisDCE. In these microcosms, the electron donor was added at day 0 and again at day 7 in some treatments (table 2). This amendment-timing experiment focused on lactate and VO in different combinations as the electron donors. The addition of WBC-2 was delayed until day 7 in all the bioaugmented treatments. Soil from the perched aquifer at site 01U611R was again used in this experiment, but one treatment included silty soil collected from the base of the aquifer at a depth of 12.5–14.5 ft bgs. Groundwater collected from 01U611R was used in this experiment, rather than simulated groundwater, to obtain biodegradation rates representative of site conditions.

The final set of microcosm experiments focused on the selection of commercial products for use in the pilot test (table 2). Many donor products that are commercially available are mixtures of fast- and slow-releasing compounds in which lactate is often used as an initial fast-release donor in the mixtures. Commercially available VO is in the form of emulsified vegetable oil (EVO) to increase distribution within the treatment zone for bioremediation. Two commercial EVO products that could be used in the field pilot test were selected for evaluation in the next experiment: SRS-SD from Terra Systems (Claymont, Delaware) and EOS Pro by EOS Remediation (Cary, North Carolina). Both products contain a fast-release donor: 5.5 percent sodium lactate in the SRS-SD and 4 percent of a glycol-based donor in EOS Pro. The slow-release components of both commercial mixtures consisted of 60-percent EVO. Both mixtures also contained vitamin B₁₂ and micronutrients to enhance microbial activity. The EVO products were compared to two commercial fast-release lactate products (table 2).

All these microcosms were incubated in an anaerobic chamber at room temperature (22 °C to 23 °C). Microcosm treatments were prepared in duplicate or triplicate for repeated aqueous sampling from the bottles, until the last sampling time point when bottles were sacrificed. Duration of the experiments varied between 22 and 63 days with 5–7 sampling points (Mejia and others, 2025). For all experiments, aqueous samples were withdrawn at each time point for analysis of VOCs and hydrocarbon gases, including methane, ethene, ethane, and acetylene. VOC samples were collected in duplicate in 4 mL glass vials with Teflon-lined septa and preserved with one drop of 1-to-1 hydrochloric acid. Water samples (2 mL each) were collected in duplicate for hydrocarbon gas analysis by injecting them into 20 mL serum bottles with butyl rubber stoppers that were pre-prepared with preservative and purged with nitrogen gas as previously described (Majcher and others, 2007; Lorah and others, 2022). For select treatments, aqueous samples were also collected at some time points for analysis of nonvolatile dissolved organic carbon, volatile fatty acids (VFAs), and major ions (refer to the “Groundwater Sampling and Geochemical Analysis” section for more information).

Bioremediation Pilot Test Design and Implementation Methods

Pilot Test Design

The limited hydrologic data available after the removal of the building infrastructure and the heterogeneous distribution of TCE in the shallow perched aquifer at Site K added complexity to designing a pilot test to evaluate bioremediation. To maximize site data, the selected pilot design utilized three separate areas of Site K for field tests and are referred to in this report as the GS1, GS2, and GS3 treatment areas or test plots (fig. 5). Field tests were conducted in each treatment area using one injection well surrounded by monitoring wells, a test design similar to a modified push-pull test (Solutions-IES, Inc., 2006). Monitoring locations were placed along presumed (and most likely downgradient and lateral) transects in each treatment area, providing more monitoring wells than typical for a push-pull test. The modified push-pull tests allowed the effects of variable field conditions, including variable groundwater flow directions and TCE distribution, to be captured and design gaps to be assessed without making the assumptions required for a larger-scale pilot test. Design parameters needed for a full-scale remedy, including the injection radius of influence (ROI), groundwater velocities, and distribution of electron donor and culture (Battelle Memorial Institute and others, 2002; Solutions-IES, Inc., 2006; Interstate Technology and Regulatory Council, 2020), were thus obtained in each treatment area.

Treatment areas include GS1 and GS2 in the plume source area near the former sump and GS3 in the downgradient, dissolved plume (fig. 5). A total of 42 wells were installed for the pilot test: one injection well and 13 monitoring wells (numbered MW1 through MW13) per treatment area (fig. 5; table 1). The name of each well indicates the treatment area in which it was installed. For example, GS1-I is the injection well of treatment area GS1.

Each area’s MW1 was installed upgradient from its respective injection well to provide a control for areas affected by amendment injections. The remaining monitoring locations could provide the distribution of amendments in three directions from the injection well. Each treatment area is approximately 30 ft wide by 60 ft long. These dimensions were calculated to enable 1 year of amendment monitoring based on the estimated ROI and groundwater velocity in the area determined from the pre-design characterization and include an adequate margin of safety, per guidance from Battelle Memorial Institute and others (2002), Solutions-IES, Inc. (2006), and standards of practice.

Performance Monitoring

Performance objectives were defined for the 1-year bioremediation pilot test at Site K to guide the monitoring approach, including the measurement of hydrologic characteristics, selection of geochemical and microbiological analyses, and sampling frequency. The performance objectives were as follows:

Hydrologic measurements were made to evaluate groundwater flow directions and characterize temporal changes in water table elevations. Geochemical constituents measured to assess the efficacy of the bioremediation test included concentrations of TCE and associated degradation products, dissolved gases, cations, anions, redox constituents, organic acids, total and dissolved organic carbon, alkalinity, and compound specific isotope analysis (CSIA) for stable carbon isotopes (table 5). Samples were also collected for microbial community analyses to monitor the movement of the injected culture. Field parameters (specific conductance, pH, and temperature) were also measured to assist in evaluating the movement of the injected amendments and the conditions for microbial activity. Hydrologic, geochemical, and microbial analytical methods are given in the following sections.

Hydrogeology

The lithology at the replacement well sites (01U608R, 01U609R, and 01U611R) that were installed near the TCE source area at Site K differed from available descriptions of the original sites (01U608, 01U609, and 01U611) at a depth of 1 to 4 ft bgs (Mejia and others, 2025). Previous lithologic descriptions noted medium to coarse sand below 1-ft-thick concrete, whereas uniformly fine to medium sand was present at the replacement well sites because of backfilling with clean sand after soil excavation (fig. 4). Fine to medium sand was present at the replacement well sites below a 0.3-ft-thick soil zone to a depth of 6.4–9.0 ft, where the gray clay that marks the aquitard began at wells 01U608R, 01U609R, and 01U611R (Mejia and others, 2025). Soil cores collected later at GS1-I, a well installed for the pilot test near well 01U611R, showed a similar depth to the top of the lower aquitard of 7.8 ft (Mejia and others, 2025). Screens (2-ft long) in the replacement wells were placed at the bottom of the perched aquifer to obtain TCE concentrations that could be affected by residual DNAPL or sorbed TCE associated with the clay at the base of the aquifer. Downgradient from the source area close to the trench, the depth to the basal aquitard was about 20 ft in soil cores at well site GS3-I installed for the pilot test (Mejia and others, 2025), confirming that the perched aquifer thickens near the trench as previously reported (fig. 2).

Site redevelopment activities did not appear to alter the relatively low horizontal head gradients in the source area, as evidenced by the similar water-level elevations among the wells near the source area in 2003–5 compared to those in 2021 (fig. 6). The effect of the excavation trench is seen in the lower water-level elevations in 01U603 and 01U615 compared to the source area, historically and in 2021 (fig. 6). Water-level data collected in the source area during the pre-design characterization indicated substantial seasonal changes in water-level elevations, decreasing by more than 2 ft in July 2021 compared to May 2021 (fig. 6), and a rapid response to precipitation (fig. 7). As a result of removal of concrete foundations and stormwater drains, the water table could show a greater response to recharge events and potentially increase the horizontal flow gradient. Historical continuous water-level measurements, however, are not available to determine if site redevelopment has altered the hydrologic response. Removal of the stormwater piping in 2014 was hypothesized to have resulted in increased recharge to the aquifer, causing a rise in the water table that was observed from 2014 through 2018 (PIKA Arcadis, 2019). The maximum historical water-level elevations in 01U608, 01U609, and 01U611, which were measured in 2014, were similar to the water levels measured in May 2021 in the source area (fig. 6).

The 2021 hydraulic head measurements and slug tests were used to calculate horizontal hydraulic conductivities (Kh) and flow velocities at the site to design the placement of injection and monitoring wells for the pilot test. Slug tests were conducted in the replacement wells 01U608R, 01U609R, and 01U611R and in the two existing wells near the trench, 01U603 and 01U615 (Mejia and others, 2025). The Kh determined from the two existing wells were substantially lower than those obtained from the replacement wells, possibly because of screen deterioration or another localized problem; thus, only slug test data from the replacement wells were used (table 9). The most probable calculated values of Kh for the perched aquifer at Site K ranged from about 5 to 9 feet per day (ft/d; table 9). Bulk density measurements of soil core samples (standard measurement of the weight of known volumes of dried soil) from the screened intervals of the replacement wells were used to obtain total porosity estimates of 47 percent. A range of effective porosity values of 0.2–0.4 was selected as indicative of heterogeneous unconsolidated material (Domenico and Schwartz, 1990) and used to calculate groundwater velocity. An effective porosity value of 0.3 was used in final design calculations.

Because of the flat head gradients measured in the source area and Building 103 footprint (approximated to be 0.0008 ft/ft), flow velocities calculated using the full range in effective porosities were consistently estimated to be less than 0.5 ft/d. A reactive zone of 15–20 ft would be needed in the pilot test to establish a hydraulic residence time of 35 days, if the flow velocity is 0.5 ft/d. Thus, the pilot bioremediation test was designed with short lateral distances between injection and monitoring points in the source area to be able to observe transport effects within a year.

Following the installation of test plot injection and monitoring wells and immediately prior to the initial biostimulation injection, baseline hydrologic conditions were quantified and are summarized and compared to the pre-design investigation in this report. The groundwater elevation, as shown in contours at each plot location (fig. 8), indicates little change (less than 0.2 ft) in elevation across test plot GS1. The direction of groundwater flow seems to be along the centerline of the monitoring network (GS1-MW1 to GS1-MW9). Outside of the excavated area nearby the GS2 test plot, however, there is more change in groundwater elevation (up to 0.5 ft). The direction of groundwater flow in the GS2 plot is not aligned with the monitoring network but instead indicates southward flow from the excavated area during the baseline measurement (fig. 8). Hydraulic conductivity in the injection wells of the source area, GS1-I and GS2-I, were similar to estimates from the pre-characterization discussed previously and ranged from about 5–6 ft/d (Mejia and others, 2025). In the GS3 test plot, located near the trench, groundwater gradients were greater than in the source area (elevation change of greater than 2 ft across the test plot). Groundwater flow in the GS3 plot is likely influenced by the pumping toward the collection trench, skewing groundwater flow slightly northward of the centerline of the plot (fig. 8). Hydraulic conductivity determined for the injection well in test plot GS3 (GS3-I) was about an order of magnitude below that of the source area (about 0.5 ft/d), consistent with variable sand and clay lenses noted in the lithologic log (Mejia and others, 2025).

Volatile Organic Compounds in Soil

Previous investigations suggested that source materials persisted below the depth of excavation and backfilling at the site, as shown in figure 4 by the historical soil concentrations exceeding the Soil Reference Values at 6-ft depth (Wenck Associates, Inc., 2016). Soil core samples analyzed for VOCs in 2020–21 showed that the highest concentrations were near or in the clay at the base of the perched aquifer (fig. 9). The highest TCE concentration was nearly 200 mg/kg, measured at a depth of 12 ft in the soil core from 01U611R. A similarly high concentration of about 130 mg/kg TCE was measured at a depth of 7.3 ft in the soil core from 01U609R, within the clay layer that began at a depth of 6.4 ft. Historical data from soil core samples in the source area also showed the highest TCE concentrations in the clay at the base of the perched aquifer (fig. 4; table 10). The deepest soil sample was collected 13–14 ft below land surface (sample RI-1013-18; fig. 4) and had the maximum concentration of 804 mg/kg of TCE (table 10). Of the historical soil samples collected, the second highest TCE concentration was 140 mg/kg (sample GP18, table 10), measured in a sample from the clay at 11 ft depth and consistent with the USGS maximum concentration measured in the soil at 01U609R (fig. 9).

The soil cores collected by the USGS verified that soil containing a high concentration of VOCs was left in place in the source area in the 2014 soil excavation and needed to be considered in the design for the bioremediation pilot test. The historical and recent USGS soil analyses showed that TCE was the VOC that occurred in the highest concentrations (fig. 4; table 10). Although cisDCE was present in most soil samples, concentrations were generally more than an order of magnitude lower than TCE concentrations. Two exceptions were the soil samples collected at 8.5 ft bgs and 9.1 ft bgs at 01U608R, where concentrations of TCE and cisDCE were similar (fig. 9). The higher proportion of cisDCE in the saturated soil at 01U608R may indicate relatively high rates of anaerobic biodegradation of TCE by reductive dechlorination at this site. The high VOC concentrations present in soil samples collected in the current study indicate that the KMnO₄ added after the sump excavation in 2009 did not effectively oxidize nor treat the soil in the entire source area, as was also indicated by the groundwater sampling following the KMnO₄ addition (Wenck Associates, Inc., 2016).

Volatile Organic Compounds and Redox Conditions in Groundwater

Groundwater VOC concentrations measured during the pre-design characterization showed that 01U609R and 01U611R had the highest concentrations, with a maximum TCE concentration of 57,700 μg/L in 01U609R (fig. 10). TCE, cisDCE, and transDCE were the predominant VOCs detected in the groundwater. TCE comprised over 75 percent of total VOC concentrations detected in the groundwater at 01U609R and 01U611R, whereas cisDCE had the second-highest concentration of the VOCs measured. Of the wells sampled in the source area, only 01U608R had cisDCE concentrations that were greater than the TCE concentrations. Thus, the groundwater data were consistent with soil sample results that indicated substantial reductive dechlorination of TCE to cisDCE at 01U608R (figs. 8, 9). The low VC concentrations in the groundwater indicated that TCE reductive dechlorination was incomplete (fig. 10). Ethene and ethane, nonchlorinated end products of reductive dechlorination of TCE, were not detected in any of the wells sampled during the pre-design characterization (Mejia and others, 2025).

Although fluctuations in concentrations have occurred, evaluation of historical data showed that groundwater VOC concentrations in the source area have remained consistently high. Groundwater from well 01U609, which was a control well during the HRC pilot test in 2000, had TCE concentrations ranging from 61,000 to 110,000 μg/L, and cisDCE concentrations ranged from 10,000 to 20,000 μg/L (Sola, 2001). The groundwater VOC concentrations in 01U609R in 2021 (57,700 μg/L TCE and 7,590 μg/L cisDCE) were consistent with these historical concentrations (fig. 10). The historical and pre-design VOC concentrations in groundwater from these two wells showed evidence of nearby residual DNAPL, based on the general rule of thumb of TCE concentrations exceeding 1 percent of the solubility (Kueper and Davies, 2009).

Concentrations measured in well 01U611R also were consistent with historical VOC concentrations measured in groundwater from well 01U611 (table 11). TCE concentrations ranged from 4,900 to 11,000 μg/L between 2009 and 2014, compared to concentrations of 3,090 and 11,700 μg/L, respectively, in 2020 and 2021 (table 11). Similar to the soil VOC results, the groundwater concentrations in the source area showed little effect from the KMnO₄ addition in 2009, or little effect from the continual operation of the trench collection and treatment system.

In addition to VOCs, groundwater samples collected by the USGS in 2020–21 were sampled for field parameters, redox-sensitive constituents, NVDOC, major ions, and trace metals (Mejia and others, 2025). Historical data for constituents other than VOCs, however, were generally limited to data collected in association with the previous bioremediation pilot tests (Sola, 2001). Recent and historical groundwater data showed near-neutral pH in the perched aquifer, which is advantageous for support of reductive dechlorination of TCE. Redox-sensitive constituents indicate that the contaminated groundwater in both the source area and downgradient was relatively reducing during the 2020–21 pre-design characterization with dissolved oxygen concentrations less than 0.5 mg/L (fig. 11), whereas oxic conditions with greater than about 3 mg/L dissolved oxygen were reported in 2000 (Sola, 2001). Methane was consistently detected in the groundwater samples collected in 2020–21, and concentrations greater than 0.40 mg/L in the source area particularly showed reducing conditions (fig. 11).

The reducing conditions in the groundwater in 2020–21 were consistent with cisDCE detected in the groundwater, which indicated at least partial reductive dechlorination of TCE occurred (fig. 10). NVDOC concentrations ranged from 2.43 to 9.22 mg/L in the six wells sampled during the pre-design characterization (Mejia and others, 2025), indicating that some natural organic carbon donor was available in the perched aquifer to support TCE reductive dechlorination. Thus, the VOC composition and concentrations of redox constituents and NVDOC during the pre-design characterization verified that an anaerobic bioremediation approach was feasible for Site K, despite the rapid response to precipitation events and subsequent potential for transport of dissolved oxygen to the shallow aquifer.

Natural Biodegradation and Effect of Bioaugmentation

Microcosm experiments were conducted with the native microbial communities obtained from monitoring wells 01U603, 01U611R, and 01U615 by combining sand from PMSs, deployed to obtain the native microbial community that prefers attachment, and water collected from the associated wells. The native microbial community was biostimulated with the addition of sodium lactate as an electron donor and compared to replicate treatments that were bioaugmented with WBC-2 (table 2). Rapid biodegradation of TCE was observed in all bioaugmented treatments, where complete degradation of TCE and all chlorinated daughter products occurred by day 7 (figs. 12, 13). In addition, the lower initial (day 0) TCE in the bioaugmented microcosms compared to those without added WBC-2 indicate that some TCE degradation occurred in the bioaugmented microcosms in the several hours between microcosm preparation and the day 0 sampling (fig. 12). The increase in TCE between day 0 and day 3 in the nonbioaugmented treatments was likely due to desorption, whereas biodegradation in the bioaugmented microcosms appeared to be more rapid than desorption of TCE from the sand matrix (fig. 12).

Production of ethene in the bioaugmented treatments verified complete reductive dechlorination of TCE (fig. 13). Because of the rapid biodegradation, little accumulation of chlorinated daughter products occurred in the bioaugmented treatments. In contrast, treatments that were not bioaugmented with WBC-2 showed removal of amended TCE only in the microcosms for well 01U611R (fig. 13). The nonbioaugmented treatments constructed with the 01U611R microbial community showed incomplete degradation of TCE, however, as evidenced by high DCE accumulation over the 21-day experiment (fig. 13). No measurable ethene was produced in the nonbioaugmented treatments.

Rapid methane production occurred in the bioaugmented microcosm treatments, reaching concentrations greater than 200 µg/L by day 3, and showed that the donor addition was sufficient to achieve and sustain TCE degradation (Mejia and others, 2025). The lack of biodegradation in the microcosms without WBC-2 indicated that the microbial community is lacking sufficient populations of OHRB. The results of these microcosms demonstrated the need for bioaugmentation to ensure complete degradation of TCE and associated daughter products. In addition, the results showed that the activity of the bioaugmented WBC-2 culture was not inhibited by the groundwater chemistry or competition with the native microbial population in the perched aquifer.

Electron Donors and Bioaugmentation Timing

In microcosms constructed with aquifer soil collected from site 01U611R and simulated groundwater, electron donors to support reductive dechlorination were evaluated, including lactate and corn syrup as fast-release compounds and whey, soy-based VO, and 3D as slow-release compounds. Each of the electron donors requires a fermentation step or steps to release hydrogen, which is utilized as the electron donor by microbial dechlorinating bacteria. Many commercial products are mixtures of fast- and slow-release compounds in which lactate is often used as an initial fast-release donor in the mixtures. Lactate treatments also provide a reference for optimum WBC-2 activity because the culture is maintained using lactate as the primary donor. Electron donors were tested in microcosms with and without added WBC-2 to evaluate the need for bioaugmentation (table 2).

Overall biodegradation of TCE and its chlorinated daughter products in each of the microcosm treatments was calculated as the change in molar concentrations of total organic chlorine in the microcosms. The most total organic chlorine removal was observed in WBC-2 bioaugmented treatments with added lactate and VO as the electron donors (fig. 14). TCE degradation was observed in each bioaugmented microcosm to which an electron donor was added (Mejia and others, 2025). However, degradation of TCE stalled at cisDCE in microcosms with added whey, corn syrup, or 3D (Mejia and others, 2025) and resulted in an overall total organic chlorine removal of less than about 40 percent in the whey- and corn-syrup-added microcosms and an increase in total organic chlorine in the 3D-added microcosm (fig. 14). In the bioaugmented microcosm with added 3D, the production of cisDCE was greater than the amount of TCE initially added and measured in the microcosm on day 1, causing an increase in the total organic chlorine that peaked on day 18 (fig. 14). This initial increase in total organic chlorine in the 3D treatment was likely due to desorption or dissolution of TCE in the site soil used in the microcosms and subsequent biodegradation to DCE (predominantly cisDCE). Similar desorption and biodegradation of TCE in the soil likely occurred in the other bioaugmented donor treatments as well, but the degradation rate was rapid enough to result in a continual overall removal of total organic chlorine (fig. 14). It is also possible that the composition of the 3D enhanced desorption from the site soil, without enhancing the biodegradation rate, compared to the other donors.

Live control treatments without added WBC-2 were done for each donor tested (table 2), but biodegradation of TCE was minor and incomplete in the live controls. As an example, figure 15 compares the VO treatments with and without added WBC-2; data for all treatments are in Mejia and others (2025). In the bioaugmented VO treatment, TCE degraded to near detection levels by day 24, as simultaneous production of DCE and VC occurred (fig. 15A). A relatively small increase in total cVOCs by day 24 above the initial concentration indicated that desorption or dissolution of residual TCE in the soil did affect the overall removal rate. However, total concentrations of cVOCs decreased substantially by the end of the microcosm experiment in the bioaugmented VO treatment, indicating complete biodegradation to nonchlorinated daughter products occurred (fig. 15A). In contrast, the nonbioaugmented VO treatment had substantial TCE remaining at the end of the experiment and showed only DCE production (fig. 15B). Total concentrations of cVOCs were increasing at the end of the nonbioaugmented treatment, indicating that biodegradation was incomplete and slower than TCE desorption rates (fig. 15B). Methane production occurred early in all the treatments (Mejia and others, 2025), indicating that donor addition (biostimulation) created sufficiently reducing conditions to support reductive dechlorination with or without the addition of WBC-2 (bioaugmentation). However, biostimulation coupled with bioaugmentation resulted in faster dechlorination rates of the TCE.

Although the bioaugmented microcosms with lactate or VO showed faster and more complete TCE degradation than the other electron donor treatments, degradation rates were slower than desired for the VO treatment (fig. 14). In addition, microcosms constructed with deeper soil from the bottom of the perched aquifer (mix of sand, silt, and clay) contained high sorbed/residual concentrations of TCE that resulted in accumulation of DCE concentrations greater than the initial amended TCE concentrations in the microcosms (Mejia and others, 2025, refer to treatment containing “DS” in name). An additional “amendment timing” experiment was conducted to evaluate if delaying the addition of WBC-2 after the electron donor reduced the accumulation of DCE and provided an overall faster degradation rate of the chlorinated ethenes, particularly for VO in combination with lactate to provide short- and long-term donors (fig. 16; table 2). In these microcosms, the electron donor was added on day 0 or days 0 and 7, and WBC-2 was added on day 7 (table 2). Delaying WBC-2 addition to the microcosms with lactate or lactate and VO as donors increased the degradation rate substantially compared to microcosms with WBC-2 added at day 0 (figs. 14, 16). The continual decrease in the total organic chlorine in the lactate and VO treatments in figure 16 showed that no stall of dechlorination occurred at the DCE production step. For the lactate treatment with WBC-2 added at day 0 (WBC2-TCE-L), the total organic chlorine removal stalled at apparent 25-percent removal between days 10 and 24 (fig. 14). In contrast, the lactate treatment with delayed addition of WBC-2 (DW-L-MS) showed a continual decrease in the total organic chlorine with more than 70-percent removal within 28 days (fig. 16). A similar rapid decrease in total organic chlorine was observed in treatments with either lactate alone (DW-L-MS) or lactate and VO as the donor (DW-L-VO-MS and DW-L-DVO-MS; fig. 16). First-order degradation rate constants for total organic chlorine were determined by an exponential fit to the total organic chlorine concentrations after WBC-2 addition at day 7; the degradation rate constant is the exponential constant (fig. 16). First-order rate constants for total organic chlorine removal in treatments with lactate or lactate and VO as the donor ranged between 0.061 and 0.047 per day, corresponding to half-lives of 11–15 days for the total organic chlorine (fig. 16). Based on the results of these experiments, lactate and VO were selected as suitable donors, combined with bioaugmentation with WBC-2, for the pilot bioremediation test.

The next microcosm experiment focused on testing commercial products containing lactate and VO that could be used in the pilot test. VO is commonly applied in the form of emulsified vegetable oil (EVO) to increase distribution within the treatment zone. Two commercial mixtures that consist of 60-percent EVO, SRS-SD from Terra Systems, and EOS Pro by EOS, were selected for evaluation (fig. 17; table 2). In addition to the EVO as a slow-release donor, both products contain a fast-release donor: sodium lactate (5.5 percent) in the SRS-SD and a glycol-based donor (4 percent) in EOS Pro. Both products also contain vitamin B₁₂ and micronutrients to enhance microbial activity. Comparable microcosm treatments were constructed with 0.2-percent EVO by mass of sediment. The SRS-SD treatment resulted in the most complete removal: more than 90 percent of the total organic chlorine was removed during the 43-day incubation (fig. 17). The Wilclear lactate treatment had a slightly faster removal than the SRS-SD but would not be expected to provide a donor source as long as the EVO. Thus, the EVO product SRS-SD was selected for use in the pilot test.

Amendment Delivery, Radius of Influence, and Groundwater Flow

Test plot GS1, nearby well 01U611R, is entirely within the previous excavation, making conditions the most lithologically homogeneous of the three test plots. The initial biostimulation injection at GS1 (October 19, 2021) took place during a time of low groundwater elevation (figs. 19A, 20). The effects and transport characteristics of the injection mixture were examined along the centerline of the test plot (tracking from GS1-MW1 to GS1-MW9; fig. 5), which was the intrinsic direction of groundwater flow. A response was measured in monitoring well 01U611R, 2.2 ft upgradient of the injection well, and in GS1-MW4, about 7 ft downgradient of the injection well (table 3) during the injection time (fig. 19A). The injection was completed entirely by gravity feed and averaged about 1.66 gallons per minute (gal/min) for a total of 300 gallons (gal; table 3). A corresponding response during the injection was not detected in continuous monitoring of GS1 (GS1-MW8 TCAAP Site K [USGS-450543093105407; U.S. Geological Survey, 2025]). This response is consistent with the concentrations of the added conservative tracer of bromide in plot GS1, which had a maximum in GS1-MW4 of 81.1 mg/L 20 days after the injection, indicating advective flow in the downgradient direction (GS1-MW4, 7 ft downgradient), but low concentrations in other wells (fig. 23). Examination of these breakthrough curves suggests a rapid peak, possibly missed at GS1-MW4, and then subsequent peaks within a reasonable distance from this well. The radius of influence along the flow path according to the time of bromide appearance and peak was 16.3 ft, relatively close to the design estimate of 15 ft (table 12; Nelson and Divine, 2005). Bioaugmentation injection followed biostimulation on December 1, 2021, using gravity feed at a calculated rate ranging from 1.76–2.00 gal/min (table 4). The lower total volume injected for bioaugmentation than biostimulation still triggered an elevated response in the GS1-MW4 well near the injection well (table 4).

Groundwater flow direction in test plot GS1 was generally along the centerline from GS1-I to GS1-MW4, GS1-MW7, and GS1-MW9 and trended slightly to the south towards GS1-MW11, during the baseline and February (Q1) monitoring events (fig. 20). Following the spring peak water levels (and mounding in GS1-MW3 and GS1-MW4) in May 2022 (Q2), flow shifted in the direction of GS1-MW11 to GS1-MW13 in July and October 2022 (fig. 20). However, horizontal hydraulic gradients were nearly flat, ranging from only 0.0025–0.0047 ft/ft. Combined with the measured hydraulic conductivity in the injection well (7.27 ft/d), estimates of specific discharge ranged from 6.8 to 12.5 feet per year (ft/yr). This specific discharge rate suggests that some distribution within the aquifer occurred during the injection event itself; the overall changes in water-level elevation throughout the year-long test reached a maximum of 3.68 ft between the February and May sampling events (fig. 19A), indicating secondary advective distribution in the direction of groundwater flow. No apparent reduction in hydraulic conductivity was observed before (October 2021) and after amendment injection (table 12).

Test plot GS2 straddles excavation and native materials and has the highest historical cVOC concentrations in the soil of the underlying aquitard (at well 01U609R), making it the most challenging location based on hydrogeology and chemistry. In addition, not much was known of this area prior to the pilot test, making the test design challenging. The initial biostimulation injection took place on October 21, 2021, using treated effluent from the site as described in table 3. Gravity feed resulted in a breakthrough of the injection mixture at the ground surface around the injection well. Therefore, pumping was conducted at a rate of 0.47–0.65 gal/min for a total of 200 gal. During the pumped injection timeframe, 1.8 ft of change in water table elevation was apparent in GS2-MW4 (about 7.5 ft away; fig. 19B) and 0.69 ft of change at GS2-MW8 (about 17 ft away; GS2-MW8 TCAAP Site K [USGS-450542093105508; U.S. Geological Survey, 2025]). Subsequent monitoring showed downgradient increases in bromide and multiple peaks over time. Using the method of Nelson and Divine (2005), the radius of influence estimated from these multiple peaks was about 60 ft, which is greater than the estimate for the design (table 12). Bioaugmentation injection followed biostimulation on November 30, 2021. Similar to GS1, an increase in water-level elevation was observed in nearby GS2-MW4 despite the smaller injection volume during bioaugmentation compared to biostimulation (tables 3, 4).

The predominant direction of groundwater flow in test plot GS2 was largely unknown at the site prior to the pilot test because there were few monitoring wells (figs. 1, 5). The general direction of groundwater flow during and immediately following pilot test injections was not along the centerline of the test plot, but instead followed a northeast to southwest general flow direction. An exception was during the period of high-water table conditions (May 2022), when the direction of groundwater flow appeared to follow the centerline of the plot (fig. 21) and changes in water-level elevations showed relatively high horizontal hydraulic gradients compared to October 2021 to April 2022 (table 13). Over 4 ft of elevation rise was observed between February and May 2022 sampling events. For the first 19 days following the injection, bromide concentrations increased in the downgradient well GS2-MW4 to a maximum concentration of 7.14 mg/L. Peak bromide concentrations were observed in GS2-MW7 during the May 2022 (Q2) sampling event, consistent with the water-level observations (figs. 21, 23). Horizontal hydraulic gradients between well pairs along the direction of flow ranged from 0.002 to 0.031 ft/ft in test plot GS2. The corresponding specific discharge ranged from 5.79 to 80.01 ft/yr (mean 29.7 ft/yr; table 12). No apparent reduction in hydraulic conductivity from the injected amendments was observed over the performance monitoring period (table 12).

Test plot GS3 is outside the source area and close to the collection trench, between the areas where two previous pilot tests (described in the “Previous Remedial Investigations and Activities” section) were attempted (fig. 4). The biostimulation injection was conducted on October 20, 2021. Similar to the other sites, the overall water level was low during injection; however, response at nearby wells with continuous monitoring was minimal (less than 0.25 ft). As in test plot GS1, the injection was completed entirely by gravity feed at an average rate of 1.16–1.35 gal/min for a total of 500 gal (table 3). Bromide breakthrough curves showed rapid time to peak concentration in surrounding wells, but unlike GS1 and GS2, concentration peaks were not observed along the expected flowline between the injection well and GS3-MW9. Rather, increases were noted in GS3-MW2 and GS3-MW3 in the first 19 days, but the bromide concentrations in GS3-MW4 did not peak until February (Q1), suggesting differing conditions or localized flow paths surrounding the injection well (fig. 23C). Bioaugmentation injection was conducted on December 2, 2021, and similar to the biostimulation injection, did not trigger an increase in groundwater elevation in GS3-MW4 (fig. 19C). The calculated radius of influence based on bromide was 12.74 ft (table 12), relatively consistent with the estimated design radius of influence of 15 ft.

Groundwater flow direction in GS3 trended from the injection well in a north-northwest direction, except during the May 2022 period of high water levels (fig. 22). Between February and May 2022, there was a maximum of more than 8 ft increase in groundwater elevation (fig. 19C; table 13). Under the high water-level conditions, some mounding was observed near wells GS3-MW7 and GS3-MW8, and groundwater flow directions were variable (fig. 22).

Indicators of Donor Movement and Redox Conditions

The lactate-EVO mixture used as a donor solution resulted in TOC concentrations of 1,190, 686.6, and 1,470 mg/L, respectively, in wells GS1-I, GS2-I, and GS3-I (Mejia and others, 2025). The injection solutions also contained about 200 mg/L of bromide, added as a conservative tracer. Thus, TOC-NVDOC and bromide concentrations were used as indicators of the movement of donor dissolved in the groundwater during the pilot test (fig. 23). Increases in specific conductance were often associated with increases in TOC-NVDOC and bromide (Mejia and others, 2025). Several redox-sensitive constituents, including Fe²⁺, sulfate, and methane (fig. 24), were used to evaluate whether the addition of the donor induced the anaerobic conditions required for reductive dechlorination of TCE.

High bromide concentrations (greater than 2 mg/L) compared to baseline concentrations (0.034 mg/L) were observed within 20 days after donor injection in at least one well in each of the three treatment plots. The highest bromide concentrations were observed in treatment areas GS1 and GS3, which had maximums of 81.1 and 28.5 mg/L, respectively, over the 1-year monitoring period (fig. 23). The number of wells with elevated bromide concentrations generally increased with time after injection in the three treatment areas, indicating a spread of the injected donor. The first elevated TOC-NVDOC concentrations also occurred within 20 days after donor injection in the three treatment areas. In each of the areas, elevated TOC-NVDOC concentrations between 80 and 100 mg/L were measured by day 6 following donor injection in one monitoring well (fig. 23). The maximum TOC-NVDOC concentrations, however, occurred after day 6 in the three treatment areas. The maximum TOC-NVDOC concentration was 351 mg/L in GS1-MW4 on day 20 and was followed by decreasing concentrations before a second peak of 265 mg/L on day 266 (Q3 event; fig. 23). In areas GS2 and GS3, the maximum TOC-NVDOC concentrations occurred around day 270 and day 200, respectively, and generally coincided with elevated bromide concentrations (fig. 23). However, elevated bromide concentrations sometimes preceded elevated TOC-NVDOC concentrations in a well, such as the maximum peaks of these constituents in GS3-MW4 (fig. 23). The bromide movement is likely representative of the fast-release lactate component of the donor mixture, whereas the increases in TOC-NVDOC are also due to dissolution and transport of the relatively slow-release EVO. Elevated NVDOC concentrations were measured in some wells up to 265–270 days after donor injection in GS1 and GS2 and through day 352 (Q4 event, October 2022) in GS3. Note that samples were not obtained in the GS2 area during the Q4 sampling event because of insufficient water (fig. 23).

Fe²⁺ concentrations increased after donor amendment in the three treatment areas, indicating that anaerobic iron-reducing conditions were enhanced (fig. 24). Baseline Fe²⁺ concentrations ranged between 0.069 and 0.560 mg/L in monitoring wells in the three treatment areas. Fe²⁺ concentrations in some monitoring wells downgradient from the injection wells increased to concentrations between 2 and 20 mg/L by 40 days after donor injection; concentrations then continued to increase through day 270 (Q3 event) in areas GS1 and GS2 and through day 350 (Q4 event) in area GS3 (fig. 24). Fe²⁺ concentrations were higher in areas GS1 and GS3 (maximum of 85.6 and 22.1 mg/L, respectively) than in area GS2 (maximum of 12.5 mg/L). The highest Fe²⁺ concentrations in area GS1 occurred in GS1-MW4 and GS1-MW7, which are sequentially downgradient from the injection well (fig. 24A). In area GS3, the highest Fe²⁺ concentrations occurred in GS3-MW2, GS3-MW3, and GS3-MW4, which surround the injection well (fig. 24C).

Sulfate concentrations showed a greater variability among the monitoring wells in each treatment area but were less than 35 mg/L in all wells throughout the monitoring period, except for a peak concentration of 54.5 mg/L in GS2-MW6 (fig. 24B). The sulfate variability is likely partially caused by variable increases in oxygenation of the shallow groundwater with recharge and rising water-level elevations. In area GS1, sulfate concentrations reached a maximum in many of the wells during the Q2 event, including in the upgradient well GS1-MW1 (fig. 24A). Sulfide also was commonly detected in the monitoring wells in all sampling events (Mejia and others, 2025), however, indicating that sulfate reduction occurred even during periods of oxygenation. In area GS1, sulfide concentrations were greatest in GS1-MW4, which also showed the lowest sulfate concentrations during most of the monitoring period (fig. 24A). Overall, sulfate concentrations were lowest in the groundwater in treatment area GS3, where the depth to the water table was greatest. Localized reductions in sulfate concentrations that were distinct from the concentration in the upgradient well in each test plot were most evident in area GS3. In wells GS3-MW2, GS3-MW3, and GS3-MW4 surrounding the injection well, sulfate concentrations decreased from about 11–15 mg/L to less than 3.0 mg/L following donor injection and indicate enhanced sulfate reduction (fig. 24C). The highest Fe²⁺ concentrations from iron reduction were also observed in wells GS3-MW2, GS3-MW3, and GS3-MW4.

The greatest increases in methane concentrations in the groundwater in the three treatment areas generally occurred later than the changes in Fe²⁺ or sulfate concentrations (fig. 24), consistent with increased development of reducing conditions. During baseline sampling, methane was present in relatively low concentrations that ranged from 62.5 to 570.5 µg/L in the three treatment areas. Substantial increases in methane concentrations that indicated enhancement of methanogenic conditions began more than 200 days following donor injection. Methane concentrations increased above 2,000 µg/L in some wells in the GS1 and GS3 treatment areas, reaching maximum concentrations above 10,000 µg/L by the end of the monitoring period (fig. 24). In contrast, methane concentrations did not change substantially in the groundwater in treatment area GS2, where concentrations were mostly below 350 µg/L throughout the monitoring period. Overall, the presence of methane indicated that conditions were conducive for reductive dechlorination in all three treatment areas, but the greater increases in methane in areas GS1 and GS3 likely reflect better distribution of the donor and WBC-2 culture amendments in these two treatment areas. Methane concentrations in GS1-MW1 and GS3-MW1, which are upgradient from their respective injection wells in treatment areas GS1 and GS3, did not increase substantially during the monitoring period (fig. 24), indicating that the enhanced methanogenic conditions were associated with the donor and WBC-2 amendments. The increase in methane concentrations in areas GS1 and GS3 after 200 days (Q2 event, May 2022) also coincided with an increase in groundwater temperature (fig. 25), which is known to increase microbial activity. WBC-2 was injected into the treatment areas between November 30 and December 2, 2021, 40–43 days after the donor amendment, when groundwater temperatures were low. Because bromide and NVDOC concentration increases indicated that the donor amendment was distributed to some wells before 200 days post-injection, the onset of increased methanogenic concentrations likely was dependent on the combination of WBC-2 distribution and increased water temperature. The WBC-2 culture does contain methanogens, and changes in the groundwater microbial communities will be discussed later in the section “Microbial Community Changes with Enhanced Biodegradation.”

Post-injection performance conditions in each treatment area were also compared to engineering performance criteria detailed by Borden (2017) for establishing conditions conducive to complete reductive dechlorination by EVO. These criteria include a sustained concentration of TOC or NVDOC greater than 20 mg/L, depletion of sulfate concentrations to less than 10 mg/L, and a pH between 6 and 8 standard units.

In GS1, TOC-NVDOC followed a similar trend to bromide, showing a sustained initial spike compared to baseline concentrations in monitoring wells GS1-MW4 and GS1-MW7 in the direction of groundwater flow from the injection well, whereas the spikes were not sustained in other wells surrounding the injection well (fig. 23). On November 15 (27 days post-injection), the TOC concentration in GS1-MW4 was 751 mg/L and was sustained above 20 mg/L in this well throughout the pilot test monitoring period. By May 2022 (Q2 event), TOC in monitoring wells GS1-MW7 and GS1-MW10 was greater than 20 mg/L, despite the low estimate of specific discharge. Other monitoring wells had intermittent concentrations above 20 mg/L that were not sustained. Evidence of conditions conducive to reductive dechlorination along the flowline was also confirmed with elevated concentrations of VFAs in May and July 2022 (Q2 and Q3 events), although the concentrations declined near the end of the monitoring year, particularly at monitoring points farthest from the injection well. The decline indicated a limitation to continued conditions supportive of reductive dechlorination.

Sulfate concentrations were elevated above 10 mg/L across the GS1 treatment area and showed the greatest declines to below 10 mg/L in areas of high TOC (figs. 23, 24). This decline most frequently occurred along the centerline of the test plot. Monitoring wells farthest from the injection well did not show declines below 10 mg/L at the end of the test, indicating that the added donor may not have reached these wells.

The buffering constituents added to the donor injection mixture may have been inadequate to counteract the production of acetic acid in treatment area GS1, as noted by a drop in pH to below 6 after 200 days in GS1-MW4 (a reduction of 2 pH units; Mejia and others, 2025). Lesser declines in pH were observed downgradient in GS1-MW7 and GS-MW8 as well (maximum reduction of approximately 1 unit).

Despite the higher specific discharge at GS2 compared to GS1, the TOC-NVDOC concentrations in area GS2 did not increase substantially, except in GS2-MW4. A peak TOC was observed at GS2-MW4 only 5 days after the injection, and a sustained increase in TOC-NVDOC concentrations occurred again in GS2-MW4 after 100 days (Q1). TOC-NVDOC concentrations in GS2-MW4 were above the 20 mg/L threshold (fig. 23B). Elevated TOC-NVDOC concentrations were also observed in GS2-MW6 and GS2-MW7, but concentrations were below the 20 mg/L threshold (8.41 and 10.8 mg/L on day 55 and 200; fig. 23B). Sulfate remained elevated across the monitoring network in plot GS2, except for GS2-MW4 where sustained elevated TOC-NVDOC concentrations occurred after day 100 (fig. 24). The geochemical data indicate an irregular distribution of the donor matrix to the subsurface in the GS2 area and limited establishment of conditions supportive of anaerobic dechlorination at the site, likely because of the lithologic heterogeneity of the soil matrix. Limited variability in pH was observed and was not likely factor in limiting enhanced biodegradation in GS2.

In test plot GS3, GS3-MW2 in the north-northwest direction of groundwater flow had TOC-NVDOC concentrations greater than the 20 mg/L threshold except for the Q2 event (fig. 23C). The maximum concentration in TOC-NVDOC was observed at GS3-MW4 during the Q2 event, but TOC-NVDOC concentrations declined again before the Q3 and Q4 sampling events, when groundwater flow direction shifted back to the north (figs. 22, 23). TOC-NVDOC concentrations also reached concentrations greater than 20 mg/L in GS3-MW7 and GS3-MW8 during the Q2 event but declined by the time of the Q3 event. Sulfate concentrations were depleted in areas of higher TOC-NVDOC, including GS3-MW2, GS-MW3, and GS3-MW4. Overall, sulfate concentrations were lower in GS3 than those in the treatment areas GS1 and GS2 (fig. 24).

Geochemical Evidence of Biodegradation of Chlorinated Volatile Organic Compounds

Groundwater concentrations of TCE and its sequential daughter products, including DCE, VC, and the nonchlorinated product ethene, were monitored throughout the pilot test to evaluate changes in reductive dechlorination of TCE that occurred with distribution of the injected amendments. The distribution of cVOCs and ethene during the baseline and four quarterly sampling events is shown in a grid with the upgradient control well (MW1) and potential downgradient monitoring well locations (MW2 to MW13) in relation to the injection well, providing spatial and temporal distributions in each treatment area (figs. 26, 27, 28). The changes in the molar percentage of cVOCs and ethene in the samples best represent the occurrence and extent of degradation reactions (figs. 26, 27, 28). The total mass concentrations of the cVOCs are also given in figures 26, 27 and 28 to show the heterogeneity in contamination in the shallow aquifer, which likely reflects the heterogeneity in TCE sources remaining in the soil (fig. 4). In addition to the effects of biodegradation, changes in the mass concentrations of the cVOCs during the pilot test may have resulted from DNAPL desorption and dissolution and from fluctuations with mixing of variable concentrations in the groundwater with rising and falling water levels and mixing with recharge. Overall higher cVOC concentrations were present in the groundwater in treatment areas GS1 and GS2, which were expected to be in the proximity of remaining TCE source areas from historical data and our pre-pilot test investigations, compared to area GS3, which is downgradient from the expected source areas and in a thicker part of the aquifer. Figures 29 and 30 show indicators of donor presence and biodegradation and the cVOC and ethene concentrations in groundwater along approximate flow paths downgradient from the injection wells in treatment areas GS1 and GS3; these figures provide greater temporal detail with time after donor injection and include quarterly and monthly sampling events. Temporal changes of the same constituents in the upgradient control wells for treatment areas GS1 and GS3 are shown in figure 31.

Not all the monitoring wells in each of the treatment areas showed evidence of enhanced biodegradation of TCE, which was expected because of the uncertainty in the distribution of the injected donor and WBC-2 amendments owing to uncertainty in groundwater flow directions and rates (figs. 26, 27, 28). Biodegradation of TCE was more evident in treatment areas GS1 and GS3 than in GS2 (figs. 26, 27, 28). In GS1, enhanced biodegradation of TCE to DCE was evident before the Q1 sampling event in wells adjacent to, or close to, the injection well, including GS1-MW3, GS1-MW4, and GS1-MW6, where the molar percentage of DCE or VC increased (fig. 26). The molar percentage of DCE continued to increase as TCE decreased to near or below the detection limit in GS1-MW4 and GS1-MW7 through the Q3 sampling event. By the Q3 event, enhanced TCE biodegradation was also evident in the next two rows of monitoring wells, where the molar percentage of TCE decreased and VC and ethene increased in GS1-MW10 and GS1-MW12. Complete reductive dechlorination of TCE to ethene appeared to be occurring in all these wells (GS1-MW3, GS1-MW4, GS1-MW7, GS1-MW10, and GS1-MW12) during the Q4 event (fig. 26). A lower extent of biodegradation was observed in GS1-MW8 and GS1-MW9 during the Q2, Q3, and Q4 events, where TCE degradation to predominantly DCE occurred. The cVOC and ethene data thus indicate movement of the amendments predominantly to the west and southwest in treatment area GS1, where relatively low head gradients made flow directions difficult to define (fig. 20).

In contrast to treatment area GS1, only two GS2 monitoring wells, GS2-MW4 and GS2-MW7, which are immediately downgradient from the injection well, showed consistent evidence of enhanced biodegradation of TCE with decreasing molar percentage of TCE and increasing DCE by the Q2 and Q3 events (fig. 27). Other TCE daughter products were largely undetected in the groundwater in GS2 by the Q3 event (fig. 27), and the wells could not be sampled in Q4 because of insufficient water. The cVOC results were consistent with the general lack of movement of the injected donor or increase in methane in GS2 (figs. 23, 24). The more heterogenous lithology and presence of clay layers in the perched aquifer, combined with the nearly flat hydraulic head gradients, likely limited the movement of water and distribution of the injections in GS2.

In treatment area GS3, enhanced biodegradation of TCE to DCE was evident in wells GS3-MW2 and GS3-MW4 (adjacent to the injection well) by the Q1 sampling event and in wells GS3-MW4 and GS3-MW7 (downgradient from the injection well) by the Q2 event (fig. 28). Ethene concentrations in these four wells showed that complete reductive dechlorination of the TCE and DCE was occurring before or during the Q3 sampling event. The groundwater at GS3-MW3 showed the earliest onset of ethene production, and ethene accounted for 40 percent of the cVOCs plus ethene by the Q2 sampling event. Ethene molar percentages continued to increase to more than 70 by the Q4 event in the three wells closest to the injection well (GS3-MW2, GS3-MW3, and GS3-MW4 in fig. 28). The lower initial TCE concentrations and absence of residual DNAPL, which could provide a continuing source of TCE and DCE, likely accounted for the greater percentage production of the nonchlorinated end product ethene in the groundwater in the GS3 treatment area. However, the areal extent of enhanced biodegradation appeared to be more limited in area GS3 than GS1. The groundwater in the two last rows of monitoring wells in area GS3 only showed minor reduction in total mass concentrations of cVOCs during the Q4 sampling event, but no consistent change in the molar compositions of the cVOCs and ethene was observed to indicate enhanced biodegradation (fig. 28). In test plot GS3, a depression in the water table elevations in Q2 indicated that flow from the injection wells toward the trench may have stalled or reversed directions in the area of the last two rows of monitoring wells, GS3-MW8 to GS3-MW13 (fig. 22).

Correlation between donor movement (peaks in NVDOC and VFA concentrations) and cVOC biodegradation with time after the donor injection were observed in treatment areas GS1 and GS3 (figs. 29, 30). Acetate and propionate, fermentation products of the lactate-EVO donor mixture, were the predominant VFAs detected, and their combined concentrations are shown as total VFAs in figures 29 and 30. In GS1, peaks in specific conductance values and NVDOC and VFAs concentrations downgradient from the injection well were generally concurrent (fig. 29). Peaks in NVDOC and VFA concentrations were also concurrent in GS3, but the relation with specific conductance was not consistent (fig. 30). The specific conductance values and NVDOC and VFAs concentrations in upgradient wells in treatment areas GS1 and GS3 remained fairly consistent with the baseline concentrations after the donor injection, verifying that the changes observed in the downgradient wells likely resulted from the donor and WBC-2 culture amendments (fig. 31). The predominance of acetate and propionate relative to lactate indicates rapid fermentation of the injected donor occurred. The maximum lactate concentration detected in GS1 and GS3 was 0.55 mg/L (Mejia and others, 2025), whereas the sum of acetate and propionate concentrations was greater than 500 mg/L and 80 mg/L, respectively, in GS1 and GS3 (figs. 29, 30). Continued fermentation of VFAs produces molecular hydrogen, the electron donor utilized by cVOC dehalorespirers (Löffler and others, 1999; Azizian and others, 2010). GS1-MW4, the monitoring well closest to the GS1 injection well, showed two peaks in concentrations of NVDOC and VFAs: one on day 127 and the second on day 266. The next two consecutive wells along the flow path, GS1-MW7 and GS1-MW10, showed consecutively later timing of maximum concentrations of NVDOC and VFAs compared to the first maximum in GS1-MW4 (fig. 29). Maximum concentrations of NVDOC and VFAs decreased in the wells downgradient from GS1-MW4, as would be expected from dispersion and degradation of the organic donor with distance from the injection well (fig. 29). In GS1, the maximum DCE concentrations and onset of ethene production occurred slightly after the peak NVDOC and VFA concentrations, possibly associated with the timing of hydrogen production from the VFAs (fig. 29) and achievement of methanogenic conditions (fig. 24A).

The onset of ethene production in the groundwater in plot GS3 generally occurred slightly after the peak NVDOC and VFA, as observed in groundwater at plot GS1, but the DCE concentrations were not as clearly correlated with the donor movement (fig. 30). The DCE concentrations in GS3 began to decrease by day 50, soon after the donor injection and immediately after the WBC-2 culture injection on day 40, whereas the maximum NVDOC and VFA concentrations were observed around day 200 or later (fig. 30). Because GS3 is not in the known TCE source area, the increase in TCE concentrations and the more extended period of increase in DCE concentrations that occurred post-injection in GS1 was not observed (figs. 29, 30).

Simultaneous production of TCE and DCE appeared to occur initially in area GS1 where movement of the injected donor was evident, and the molar concentrations of DCE produced were greater than the TCE concentrations in wells GS1-MW4 and GS1-MW10 (fig. 29). The relatively high DCE production likely indicates the simultaneous desorption or dissolution and biodegradation of residual TCE in the soil in this source area (fig. 29). In treatment area GS3, where the concentrations of TCE and DCE were substantially lower by comparison, the molar concentrations of ethene sometimes increased above the combined concentrations of TCE, DCE, and VC in the groundwater, indicating rapid degradation of the cVOCs. Although historical soil excavation data did not show residual TCE in or close to the soil in the GS3 treatment area (fig. 4), it is possible that desorption and degradation of TCE and DCE in clayey layers in the GS3 treatment area resulted in the high ethene. Ethene concentrations in the groundwater in GS3 were often greater than VC concentrations, indicating rapid degradation of VC (fig. 30). Ethene concentrations above 0.36 micromoles per liter (µmol/L; or 10 µg/L) has been noted as evidence that complete dechlorination of chlorinated organics occurred (U.S. Environmental Protection Agency, 1998), and the ethene concentrations produced in the groundwater in GS1 and GS3 treatment areas greatly exceeded this criterion (figs. 29, 30).

In all the monitoring wells, the predominant DCE daughter compound measured was cisDCE, followed by transDCE. The 11DCE occurred less frequently and at lower concentrations than cisDCE or transDCE (Mejia and others, 2025). The mean ratios of cisDCE to transDCE were typically greater than 10 in treatment areas GS1 and GS3 through the four quarterly sampling events, whereas the mean ratios of cisDCE to 11DCE were typically greater than 100. Ethane, another potential nonchlorinated daughter product of TCE, was detected in addition to ethene but infrequently, with only 6 detections that ranged in concentration from 7.19 to 20.9 µg/L during the pilot test (Mejia and others, 2025). Acetylene, a potential product of abiotic reduction of TCE (Ferrey and others, 2004), was not detected in any of the groundwater samples (less than 6 to less than 37 µg/L, Mejia and others, 2025).

Carbon Isotope Analyses as Evidence of Biodegradation of Chlorinated Volatile Organic Compounds

CSIA of TCE, cisDCE, transDCE and VC were measured to provide another line of evidence of the occurrence and extent of anaerobic biodegradation (fig. 32). Groundwater samples for CSIA were collected during the Q4 event where evidence for biodegradation of cVOC was observed (figs. 29, 30), as well as the upgradient wells in each treatment area (GS1-MW1 and GS3-MW1; fig. 31). Replicate samples analyzed for GS3-MW3 showed excellent agreement (fig. 32B). As microbes preferentially degrade TCE containing the lighter ¹²C isotopes before those containing ¹³C, the remaining dissolved TCE in groundwater will become isotopically heavier as biodegradation continues, resulting in more positive (and less negative) δ¹³C values (Hunkeler and others, 2008). Undegraded, pure-phase TCE has been reported to be isotopically light, having δ¹³C ranging from −34 to −22 per mil (‰) with a mean of −28‰ (Mundle and others, 2012). TCE carbon isotope ratios that are heavier than this range for pure-phase TCE would indicate that the TCE has been biodegraded. The TCE daughter compounds are initially lighter in ¹³C than the parent TCE because of the preferential incorporation of ¹²C into the daughter compounds. As the daughter compounds themselves are then biodegraded, however, each daughter compound can become isotopically heavier, causing their δ¹³C values to become greater than the initial TCE δ¹³C (Song and others, 2002; Mundle and others, 2012; Phillips and others, 2022). If the δ¹³C of daughter compounds increases by more than 2‰ compared to the parent compound (TCE), it is considered to indicate active biodegradation (Hunkeler and others, 2008). Where residual TCE source existed, continued dissolution of TCE, without biodegradation, would contribute isotopically lighter TCE into the groundwater and counteract the trend toward isotopically heavier TCE that would result from biodegradation. Despite this complicating factor, the δ¹³C of daughter compounds along the flow paths in GS1 and GS3 indicated that biodegradation was occurring in several of the wells downgradient from the injection wells (fig. 32).

In treatment area GS1, the δ¹³C of TCE in the upgradient well was −25.5‰, which is within the range of pure-phase TCE and remained about the same in the next two wells downgradient from the injection well, GS1-MW4 and GS1-MW7 (fig. 32A). The δ¹³C of TCE increased to −20.1‰ in the next well along the flow path, GS1-MW10, indicating that biodegradation of TCE was outweighing the effect of residual TCE dissolution on the carbon isotope values. In addition, the δ¹³C of cisDCE, transDCE, and VC indicated that biodegradation was occurring, as each daughter compound was isotopically heavier in GS1-MW4, GS1-MW7, and GS1-MW10 compared to the upgradient well GS1-MW1 (fig. 32A). In contrast, the δ¹³C values of TCE and its daughter products in GS1-MW13 were similar to those in the upgradient well, supporting the conclusion reached from the cVOC and ethene composition that the injected amendments had not reached this well nor enhanced biodegradation (fig. 26). Trends in the δ¹³C values in the groundwater of the GS3 treatment area were similar to those in area GS1. The δ¹³C of TCE, cisDCE, and VC were higher (isotopically heavier) in the downgradient wells GS3-MW3 and GS3-MW4 than in the upgradient well GS3-MW1, showing that biodegradation occurred (fig. 32B). Further downgradient in GS3-MW10, the δ¹³C of TCE and its daughter products were similar as those in the upgradient well GS3-MW1, indicating that the enhanced biodegradation with the injected amendments did not reach this well as of the Q4 event.

Field Biodegradation Rates of Chlorinated Volatile Organic Compounds

First-order biodegradation rates were calculated using DCE concentrations over time in one well in treatment area GS1 and two wells in GS3, each downgradient from the injection well (table 14). In area GS1, the GS1-MW10 that was selected for rate calculations may have been affected by dissolution and degradation of residual TCE sources in the soil, causing the initial high production of DCE (fig. 29B). However, CSIA results indicated that TCE and DCE in GS1-MW10 were less affected by residual DNAPL sources than GS1-MW4 or GS1-MW7 along this flow path (fig. 32). Ongoing DCE production from residual source that could not be accounted for in the rate calculations would result in the DCE degradation rate to appear lower than the actual rate. Ethene production rates were also calculated to provide an estimate of the anaerobic VC degradation rate. Molar VC concentrations were often lower than ethene concentrations and did not always decrease as ethene increased. DCE degradation resulted in half-lives of 36.9–75.3 days in the three wells, showing fairly consistent degradation rates in treatment areas GS1 and GS3. Ethene production resulted in half-lives of 9.48–38.5 days, indicating that VC degradation was more rapid than DCE degradation. The ethene production rates provide the best measurement of the rate of complete biodegradation of the TCE and DCE, although ethene biodegradation and abiotic losses, such as volatilization, could result in underestimated rates. Notably, the ethene production rates calculated from the field tests in areas GS1 and GS3 (table 14) were similar to the first-order rates of total organic chlorine removal calculated from WBC-2-bioaugmented microcosms with site soil amended with either lactate or VO as donors—rate constants of 0.061 and 0.047 per day, corresponding to half-lives of 11–15 days (fig. 16). Ethene can biodegrade under anaerobic conditions to ethane, carbon dioxide, and methane. Ethane was detected infrequently in the groundwater during the pilot bioremediation test (Mejia and others, 2025). Although methane was commonly detected, it can be produced through multiple metabolic processes in addition to ethene biodegradation.

Microbial Community Changes with Enhanced Biodegradation

The relative abundances of potentially dechlorinating genera, along with those frequently occurring in reductively dehalogenating co-cultures, were examined in the microbial communities from groundwater samples collected during the bioremediation pilot test (figs. 33, 34, 35). These genera are defined as those previously reported as dechlorinating or being linked to the degradation of chlorinated organic compounds (table 15). It should be noted that only select wells close to the injection wells were sampled during the biostimulation and post-bioaugmentation sampling events; microbial data are lacking for other sampling events, either because of dry wells or difficulties obtaining the large sample volumes needed for filtering, or because fewer than 10,000 sequences were obtained, indicative of a quality control failure for that sample.

Briefly, the potential dechlorinating genera varied among the treatment plots and the monitoring wells within the plots under the conditions present prior to biostimulation and bioaugmentation as well as after the injections. However, the composition of the potential dechlorinating genera in wells immediately downgradient from the injection wells (GS1-MW4, GS2-MW4, and GS3-MW4) consistently shifted to communities predominantly composed of Desulfuromonas and Geobacter after the donor amendment reached the wells. The effect of WBC-2 bioaugmentation was also apparent; increases in the relative abundances of Dehalogenimonas and Dehalococcoides were observed coincident with the onset of ethene production in some wells. Notably, the relative abundance of dechlorinating genera known to produce ethene remained low, demonstrating the importance of geochemical analyses along with microbial community analyses as indicators of bioremediation performance. The changes in microbial community structure over the course of the first year of the pilot study are detailed in the following paragraphs.

In the upgradient well GS1-MW1, potentially dechlorinating genera comprised 39.7 percent of the microbial community prior to biostimulation; of this, 34.4 percent of the sequences recovered were classified within the Desulfuromonas genus, 2.1 percent as Pseudomonas, and 1.9 percent as Geobacter (fig. 33). The electron donor arrived at GS1-MW1 within 6 days of the biostimulation injection, as evidenced by TOC measurements, but TOC levels decreased to background by day 42 and remained low throughout the monitoring year (fig. 31). The short-lived TOC increase, however, appeared to have a dramatic effect on community structure. Sampling of GS1-MW1 approximately 40 days after biostimulation indicated that potentially dechlorinating genera decreased to 3.5 percent of the sequences recovered. The methanotrophic genus Methylomonas (Findlay and others, 2016) comprised 1.2 percent of the recovered sequences, while Desulfuromonas and unclassified Dehalococcoidia comprised an additional 1.0 and 0.6 percent, respectively. The presence of an aerobic, potentially TCE-degrading methanotroph such as Methylomonas in the same sample as strict anaerobes (Desulfuromonas and Dehalococcoides) suggests the potential for redox zonation at the water table-vadose zone interface and the cooccurrence of aerobic and anaerobic cVOC degradation pathways. A generally similar profile was observed during the post-bioaugmentation sampling. Potentially dechlorinating genera comprised 7.7 percent of the sequences recovered; of these, Methylomonas comprised 6.2 percent of the total sequences recovered, and Desulfuromonas and unclassified Dehalococcoidia contributed an additional 0.5 and 0.4 percent. During the Q2 event, potential dechlorinating genera comprised 3 percent of the recovered sequence; of these, unclassified Dehalococcoidia comprised 2.4 percent of the recovered sequences. During the Q3 event, potential dechlorinating genera comprised 2.7 percent of the recovered sequences, of which 1.1 percent was classified as Dehalogenimonas and 0.7 percent was classified as unclassified members of the Dehalococcoidia. Potential dechlorinating genera comprised 1.4 percent of the recovered sequences during the Q4 event, of which unclassified members of the Dehalococcoidia comprised 0.8 percent. Sequencing results from the Q1 event for this well did not pass quality control. The effect of the early TOC peak on the microbial community in GS1-MW1 did not seem to enhance production of TCE daughter products (fig. 31) compared to wells downgradient from the injection well, where elevated TOC-NVDOC concentrations occurred later (after bioaugmentation) and were sustained longer than in GS1-MW1 (fig. 30).

Broadly similar results were observed elsewhere in test plot GS1 in the wells downgradient from the injection well and perpendicular to groundwater flow (GS1-MW5, GS1-MW2, GS1-MW3, and GS1-MW6; fig. 5). Wells GS1-MW2 and GS1-MW5 contained a similar proportion of potential dechlorinating genera (33.2 and 35.3 percent, respectively) during the baseline sampling, and Pseudomonas and Desulfuromonas were the predominant genera (fig. 33). Similar to the observations from upgradient well GS1-MW1, the relative abundance of sequences associated with potential dechlorinating genera dropped during the immediate post-bioaugmentation sampling. During the post-biostimulation sampling event in GS1-MW2, potential dechlorinating genera comprised 7.4 percent of the recovered sequences, of which Methylomonas comprised 3.8 percent, and Dechloromonas and unclassified Dehalococcoidia each comprised 1.1 percent. Potential dechlorinating genera comprised a slightly larger proportion of the recovered sequences (8.5 percent) during the Q1 event, during which the relative abundances of Desulfuromonas and unclassified Dehalococcoidia increased 2.4 percent each relative to the post-biostimulation sampling event. Although the overall proportion of potential dechlorinating genera in GS1-MW2 decreased slightly during the Q3 event to 6.9 percent, the distribution of the community shifted: Dehalogenimonas comprised 4.4 percent of the sequences, and unclassified Dehalococcoidia comprised 1.4 percent. Notably, Dehalogenimonas was not observed to have a relative abundance greater than 1 percent prior to this sampling event, suggesting the arrival of members of the WBC-2 consortium from the bioaugmentation injection. The increase in Dehalogenimonas in GS1-MW2 during the Q3 event coincides with decreasing cVOC concentrations and increasing production of ethene, indicating enhanced reductive dechlorination (fig. 30). The proportion of Dehalogenimonas in GS1-MW2 remained above 1 percent as ethene concentrations continued to increase by the Q4 sampling event (figs. 30, 33).

To the south of the GS1 injection well, the microbial community structure of GS1-MW3 followed similar trends to GS1-MW2. During the baseline sampling, potential dechlorinating genera comprised 28.9 percent of the recovered sequences, and Desulfuromonas and Pseudomonas were predominant (15.2 and 11 percent, respectively). As seen in the other GS1 wells, the relative abundance of potential dechlorinating genera decreased in GS1-MW3 during the post-biostimulation event, and such genera comprised 3.7 percent of the total sequences recovered. During the post-biostimulation event, only unclassified Dehalococcoidia comprised more than 1 percent of the sequences recovered (1.3  percent). Little change was observed between the pre- and post-bioaugmentation sampling. By the Q1 event, however, potential dechlorinating genera in GS1-MW3 comprised a larger proportion of the recovered sequences (4.3 percent): Methylomonas comprised 1.7 percent of the total sequences recovered, and unclassified Dehalococcoidia comprised 1 percent. The relative abundance of dechlorinating genera then decreased during the Q2 and Q3 sampling events before a notable shift in the community composition was observed during the Q4 event (fig. 33), coinciding with the first detections of ethene in this well (fig. 26). Potential dechlorinating genera comprised 6.3 percent of the recovered sequences during the Q4 event in GS1-MW3. The distribution of these sequences was notably different from that in the earlier time points: Desulfuromonas comprised 2.8 percent, Geobacter comprised 1.6 percent, and unclassified Dehalococcoidia comprised 0.9 percent of the recovered sequences. The decline in Desulfuromonas and Geobacter may be linked to the establishment of the highly reducing conditions required for the dechlorination of VC to ethene.

The well immediately downgradient from the injection well, GS1-MW4, had distinct shifts in microbial community composition following amendment injection. During the baseline sampling, potentially dechlorinating genera comprised 26.8 percent of the recovered sequences. Desulfurobacter and Geobacter comprised the majority of these sequences, 15.6 and 9 percent, respectively, and unclassified Dehalococcoidia accounted for an additional 1.3 percent. In contrast to the decrease in the relative abundance of dehalogenating genera observed in the other GS1 wells following the baseline sampling, the proportion of these genera increased following biostimulation to 48.5 percent of the total sequences recovered. The relative abundance of Desulfuromonas increased to 31 percent of the recovered sequences, Geobacter increased to 13 percent of the recovered sequences, and the proportion of unclassified Dehalococcoidia remained stable at 1.3 percent. The proportion of potential dechlorinating genera increased to 53.1 percent of the recovered sequences following bioaugmentation; Desulfuromonas accounted for 37 percent and Geobacter accounted for 15 percent of the recovered sequences, and unclassified Dehalococcoidia represented less than 0.5 percent of the recovered sequences. During the Q1 event, the relative abundance of potentially dechlorinating genera decreased to 25.9 percent of the recovered sequences with 14 percent of these belonging to Desulfuromonas and 11 percent to Geobacter. As shown in figure 24, VFA concentrations peaked and DCE production increased during the Q1 event, suggesting that microbial dechlorination was supported by fermentation of the injected carbon sources. The relative abundance of both Desulfuromonas and Geobacter increased dramatically by the Q2 event, to 41 and 26 percent of the recovered sequences, respectively. No other genera were observed to have a relative abundance above 0.5 percent. Desulfuromonas and Geobacter have been shown to couple acetate oxidation to the dechlorination of TCE to DCE (Sung and others, 2003, 2006), and the high concentration of DCE seen during the Q2 event suggests that this may have been the dominant process occurring at this point. The relative abundance of potential dechlorinating genera decreased sharply during the Q3 event, and Geobacter and Desulfuromonas also decreased, comprising 4.9 and 4.7 percent of the recovered sequences, respectively. The production of VC and ethene began between the Q2 and Q3 events, in tandem with the onset of methanogenesis and an increase in the relative proportion of sequences representing unclassified Dehalococcoidia to 0.1 percent of the recovered sequences. However, the sequencing results for the Q4 sampling event did not pass quality control. Microbial community data also were not available for GS1-MW7, immediately downgradient from GS1-MW4, after the baseline sampling.

In the last two rows of monitoring wells in test plot GS1, microbial community data are only available for the baseline and Q4 sampling events (fig. 33). Of the wells in these last rows, GS1-MW10 and GS1-MW12 showed the greatest evidence of enhanced biodegradation, beginning during the Q3 and Q4 sampling events (figs. 26, 29). Well GS1-MW8, which lies directly downgradient from GS1-MW7, also showed evidence of enhanced dechlorination by the Q4 event, when ethene was first detected at a low concentration compared to GS1-MW10 and GS1-MW12 (fig. 26). During the baseline sampling, potential dechlorinating genera comprised 7 percent of the sequences recovered in GS1-MW8, of which Pseudomonas, unclassified Burkholderiales, and unclassified Dehalococcoidia comprising 4.1, 1.2, and 0.8 percent, respectively. The potential dechlorinating genera in GS1-MW8 had shifted by the Q4 event, when enhanced reductive dechlorination of the cVOCs was evident; Geobacter and Desulfuromonas were predominant, accounting for 2.7 and 2.2 percent of the recovered sequences, respectively, and Dehalogenimonas and Dehalococcoides accounted for an additional 0.2 percent each (fig. 33).

Similarly, in GS1-MW10, the potentially dechlorinating community had shifted by the Q4 event to predominantly Geobacter and Desulfuromonas (2.2 and 1.5 percent of the recovered sequences, respectively) and relatively low abundance of Dehalococcoides and Dehalogenimonas (0.2 and 0.06 percent, respectively; fig. 33). Despite these relatively low abundances of potential dechlorinating genera, enhanced biodegradation of TCE to ethene was observed in GS1-MW10 by the Q4 event and indicated an important role for the relatively small proportion of dechlorinating genera. Further downgradient in GS1-MW12, potentially dechlorinating genera comprised 4.9 percent of the recovered sequences in the baseline sampling: Pseudomonas comprised 2.9 percent; Methylomonas, 1.1 percent; and Desulfuromonas, 0.9 percent. During the Q4 sampling of this well, the community shifted so that Geobacter and Desulfuromonas were predominant (4.5 and 4.3 percent of the recovered sequences, respectively) as observed in wells GS1-MW8 and GS1-MW10 (fig. 33). Despite the low relative abundance of Dehalococcoides (<0.1 percent) in GS1-MW12, dechlorination of cVOCs to ethene was first observed by the Q4 event, suggesting a role for other dehalogenating genera capable of the final dechlorination step. Dehalogenimonas species are known to be able to dechlorinate cVOCs to ethene (Moe and others, 2009) and were found at a relatively low abundance of 0.3 percent.

During the baseline sampling of the upgradient well GS2-MW1 in plot GS2, potential dechlorinating taxa comprised 5.9 percent of the sequences recovered, the majority of which were Pseudomonas and unclassified Burkholderiales (3.6 and 1.8 percent, respectively). Unclassified Dehalococcoidia comprised an additional 0.3 percent of the recovered sequences. The community structure was relatively stable throughout the pre- and post-bioaugmentation sampling and the Q3 sampling event. Potentially dechlorinating genera comprised an average of 2.8 percent of the recovered sequences. In contrast to the changes seen across the downgradient wells in plot GS1, microbial community changes in plot GS2 were greatest in GS2-MW4 (fig. 34), which was consistent with elevated TOC and dissolved organic carbon concentrations (fig. 23B) and changes in cVOCs also most evident in GS2-MW4 (fig. 27). During the baseline sampling, the most abundant potential dechlorinating genera in GS2-MW4 included Pseudomonas (6.2 percent), Desulfuromonas (3.7 percent), and Geobacter (1.3 percent). During the pre-bioaugmentation sampling, approximately 39 days after biostimulation, Desulfuromonas was observed to be the dominant genus in the microbial community, comprising 81 percent of the sequences recovered, and Geobacter comprised an additional 1.4 percent. The community structure shifted during the Q1 event; Desulfuromonas decreased to 52 percent of the recovered sequences, and Geobacter increased to 15 percent, concurrent with an increase in iron reduction (fig. 24B). The community was stable during the Q2 event: Desulfuromonas and Geobacter comprised 56 percent and 8.6 percent of the recovered sequences, respectively, and Sulfurospirillum comprised an additional 1.3 percent. During the Q3 event, the relative abundance of Desulfuromonas decreased to 19 percent, while that of Geobacter decreased to 2.3 percent, indicating a community response to the arrival of the amendment (fig. 23). The relative abundance of Sulfurospirillum remained stable at 1.8 percent of the recovered sequences.

Compared to GS2-MW4, potentially dechlorinating genera comprised a small portion of the microbial community in the wells parallel to the injection well in plot GS2, including GS2-MW5, GS2-MW2, GS2-MW3, and GS2-MW6 (fig. 34). Only baseline and pre-bioaugmentation samples were available for GS2-MW5 and GS2-MW6, and Pseudomonas was the predominant dechlorinating genus in both wells (5.8 and 4.8 percent of the total sequences recovered, respectively). During the post-biostimulation sampling, unclassified Dehalococcoidia represented 1.8 percent of the total sequences recovered in GS2-MW6; no other potentially dechlorinating genera were found to have a relative abundance greater than 1 percent of the recovered sequences.

Markedly different baseline communities were observed in wells GS2-MW2 and GS2-MW3, immediately adjacent to the injection well. Potentially dechlorinating genera comprised 12.3 percent of the recovered sequences from GS2-MW2 and 1.1 percent of those in GS2-MW3 during the baseline sampling. Potential dechlorinating genera never comprised more than 2 percent of the sequences recovered from GS2-MW3 throughout the Q1, Q2, and Q3 events. In contrast, the relative abundance of potential dechlorinating genera in GS2-MW2 remained below 5 percent of the total sequences recovered until the Q3 event, when the relative abundances of Dehalogenimonas reached 9.1 percent; Geobacter, 2.4 percent; and Desulfuromonas, 2.2 percent. The sudden increase in Dehalogenimonas in GS2-MW2 suggests the arrival of the WBC-2 culture from the bioaugmentation injection, although cVOC concentrations remained approximately the same during the Q2 and Q3 events (fig. 27).

Site GS3 is located further downgradient from test plots GS1 and GS2 (fig. 5) and has a distinct cVOC composition dominated by DCE during the baseline sampling (fig. 28). The microbial community composition at the upgradient well GS3-MW1 also was distinct in comparison to the other plots during the baseline sampling event (figs. 33, 34, 35). Unclassified Dehalococcoidia comprised 11.7 percent of the recovered sequences from GS3-MW1, but there were smaller contributions from Desulfuromonas and Geobacter, 1.6 and 1 percent, respectively. During the post-bioaugmentation sampling, unclassified Dehalococcoidia comprised 7 percent of the dechlorinating genera and remained the most abundant potential dechlorinating genus during the Q2 and Q3 sampling events (fig. 35).

Changes in cVOC composition and ethene production indicated enhanced reductive dechlorination in wells GS3-MW2 and GS3-MW3 adjacent to the injection well and wells GS3-MW4 and GS3-MW7 immediately downgradient from the injection well; these changes were evident by the Q2 sampling event in GS3-MW3 and Q3 in the other wells (figs. 28, 30). The effects of biostimulation on the microbial community were pronounced at GS3-MW3 (fig. 35). Potential dechlorinating genera were a small portion of the microbial community during the baseline sampling: Pseudomonas comprised 3.1 percent of the recovered sequences, but no other observed dechlorinating genera had a relative abundance exceeding 1 percent. Coinciding with the arrival of the donor amendment that was observed during the post-biostimulation sampling event (fig. 23C), the community composition in GS3-MW3 was dominated by Desulfuromonas, which had a relative abundance of 78.3 percent of the recovered sequences. A small additional contribution came from Geobacter, which had a relative abundance of 1.6 percent of the recovered sequences. The relative abundance of Geobacter increased to 11.2 percent by the Q2 sampling event, while that of Desulfuromonas decreased to 14.3 percent. The relative abundance of Dehalococcoides and Dehalogenimonas remained low (0.4 percent for each genus) in GS3-MW3 during the Q2 and Q3 sampling events, although dechlorination of DCE and VC to ethene and a decrease in total cVOC was observed by Q2 (figs. 28, 35).

In contrast to GS-MW3, the relative abundance of Dehalococcoides and Dehalogenimonas in GS3-MW2 showed a more prominent shift, although not until the Q4 event when data were not available for GS-MW3 (fig. 35). During the Q3 sampling event, no potential dehalogenators in GS3-MW2 exceeded a relative abundance of 1 percent. However, Dehalogenimonas, Geobacter, and Dehalococcoides comprised 13, 2.9, and 1.8 percent of the recovered sequences, respectively, in GS3-MW2 during the Q4 event. This shift in community structure, along with a shift to a VOC composition dominated by VC and ethene, indicates the arrival of organisms from the WBC-2 bioaugmentation.

In GS3-MW4 downgradient from the injection well, the composition of the potentially dehalogenating community immediately after bioaugmentation was generally similar to the community observed in GS3-MW3: Desulfuromonas represented 54 percent of the recovered sequences, and Geobacter comprised an additional 0.9 percent (fig. 35). By the Q3 event, the relative abundance of Desulfuromonas had decreased to 11.3 percent of the recovered sequences, and Geobacter had increased to 2 percent. The composition of the microbial community in GS3-MW4 was greatly different during the Q4 event, possibly because of electron donor depletion (fig. 30); no potentially dechlorinating genera had a relative abundance exceeding 1 percent, although Dehalococcoides comprised 0.7 percent of the recovered sequences, and Geobacter, 0.6 percent; and Desulfuromonas, 0.6 percent. Despite the low abundance of known dechlorinating genera, ethene concentrations continued to increase through the Q4 sampling event and the last monthly sampling event (fig. 30). Similarly, in GS3-MW7 downgradient from GS3-MW4, ethene was a substantial component of the VOC composition during the Q3 and Q4 sampling events (figs. 28, 30), and the proportion of the microbial community comprised of potential dehalogenating genera was relatively low (3.4 percent) by the Q4 sampling event (fig. 35). Dehalococcoidia and Dehalogenimonas comprised only a small portion of the community (0.5 and 0.05 percent, respectively) in GS3-MW7 by the Q4 event.

To summarize, the potential dechlorinating genera present in the aquifer under baseline conditions and in the upgradient monitoring wells varied among the three treatment plots (figs. 33, 34, 35). In contrast, the composition of the dechlorinating genera in wells immediately downgradient from the injection wells (GS1-MW4, GS2-MW4, and GS3-MW4) in each test plot consistently shifted to predominance of Desulfuromonas and Geobacter after the donor amendment had reached the wells. Distinct increases in the relative abundance of Dehalogenimonas and Dehalococcoides also were coincident with the onset of ethene production in some wells, indicating the WBC-2 bioaugmentation’s effect on the organisms. These patterns are supportive of the successful bioremediation performance at the site. The detected abundances of these dechlorinating genera that are known to enable ethene production were often relatively low; thus, the microbial community analyses did not indicate bioremediation performance at this site as strongly as the geochemical data, including the cVOCs, ethene, and stable carbon isotope data. Partly, the microbial community data were limited by missing data for some key points, spatially or temporally. The method of microbial sample collection by filtration of the well water samples may also have been a factor in the relatively low abundances of dechlorinating genera that were measured as the distribution of free-living versus soil-bound dechlorinators is not well understood. A passive method of sample collection in the wells or collection of soil cores could improve microbial community analysis. Furthermore, we anticipate that improvements in DNA extraction and sequencing technology will improve the rate of success for low-biomass samples.