Scientific Investigations Report 2006-5056
U.S. GEOLOGICAL SURVEY
Scientific Investigations Report 2006-5056
In the previous USGS evaluation of natural attenuation at OU-1 (Dinicola and others, 2002), biodegradation was determined to be responsible for significantly reducing the contaminant mass at OU-1, although natural attenuation overall was not effective enough in 2000 to meet quantitative remediation goals. Those goals were not met because of the relatively short distance between the landfill and the adjacent marsh, and because of extremely high chloroethene concentrations at a few locations beneath the landfill.
For this report, the 2001–04 data were examined for evidence of continued biodegradation of chloroethenes. Geochemical data and redox conditions were examined to determine if any changes occurred that could result in either more or less efficient biodegradation. Changes in absolute and relative concentrations of contaminants were examined as direct evidence of continued biodegradation. The rate at which the contaminant mass at OU-1 was degraded in ground water and the rate at which contaminants were discharged to surface water during 2004 were calculated and compared to rates previously calculated using 1999–2000 data.
The June 2004 and previous geochemical data collected by USGS at OU-1 (Dinicola and others, 2002; Dinicola, 2003; 2004; Dinicola and Huffman, 2004) are shown in table 6 (at back of report). The predominant redox conditions shown in table 6 were inferred using all geochemical data collected and best professional judgment. For convenience in following the discussion, the sampled wells and piezometers are grouped by location and aquifer. “Upgradient” sites are the two upper aquifer wells and one intermediate aquifer well upgradient from the landfill. “Northern plantation” and “southern plantation” sites are all upper aquifer wells and piezometers in or near the respective phytoremediation plantations. “Intermediate aquifer” sites are all intermediate aquifer wells downgradient from the landfill. No intermediate aquifer wells are in the footprint of the former landfill.
At the upgradient wells within the upper aquifer (wells MW1 3 and MW1-20), predominant redox conditions from 1998 to 2004 varied between aerobic, iron-reducing, and sulfate-reducing (table 6). Dissolved organic carbon concentrations consistently were less than 2 mg/L, and methane concentrations were consistently less than 0.3 mg/L. Although these wells are upgradient from the landfill, they are downgradient from the military base industrial and office areas and are near stormwater swales; therefore, upper-aquifer water flowing into OU-1 is not pristine. In the upgradient well in the intermediate aquifer (MW1-33), redox conditions were consistently aerobic. Well MW1-33 is in a forested area downgradient from low density off-base residential areas and does not appear to be influenced by local land use.
At the northern plantation sites, predominant redox conditions in shallow ground water were consistently anaerobic during 1996–2004 (table 6). The specific redox conditions ranged from iron-reducing to sulfate-reducing, although the widespread detection of methane (0.22–7.4 mg/L during 2004) indicated methanogenic conditions also were present. Dissolved organic carbon concentrations in the northern part of the landfill (6.7–27 mg/L in 2004) were consistently greater than concentrations measured in upgradient wells, indicating that the landfill is a source of organic substrate essential for reductive dechlorination. Throughout the northern plantation and vicinity, no consistent trends developed in predominant redox conditions or in most redox related geochemical concentrations since 1996.
At the southern plantation sites, predominant redox conditions in shallow ground water also were consistently anaerobic during 1996–2004 (table 6). The specific redox conditions ranged from iron/manganese-reducing to methanogenic, although mildly reducing conditions were more common than strongly reducing conditions. Dissolved organic carbon concentrations in the southern part of the landfill (2.7–36 mg/L during 2004) also were consistently greater than concentrations measured in upgradient wells. Similar to the northern plantation, no consistent trends developed in predominant redox conditions or in most redox related geochemical concentrations since 1996 throughout the southern plantation.
At the intermediate aquifer sites, predominant redox conditions were consistently anaerobic during 1996–2004 (table 6). For intermediate aquifer wells at the downgradient margin of the landfill (1MW-4, MW1-25, MW1-28, and MW1-29), the specific redox conditions ranged from iron/manganese-reducing to methanogenic with iron-reducing conditions measured most frequently. For intermediate aquifer wells northwest of the tide flats (MW1 38 and MW1-39), the specific redox conditions ranged from iron reducing to methanogenic. Redox conditions at the only contaminated well northwest of the tide flats (MW1-39) were predominantly sulfate-reducing. Dissolved organic carbon concentrations in the intermediate aquifer downgradient from the landfill (2.4–15 mg/L during 2004) were consistently greater than concentrations measured in the upgradient well MW1 33. Similar to the upper aquifer, no consistent trends in redox conditions or in most redox-related geochemical concentrations developed since 1996 in intermediate aquifer wells downgradient from the landfill.
Overall, no widespread changes in geochemical data and redox conditions occurred since 1996 at OU-1 that could result in either more or less efficient biodegradation. Redox conditions varied substantially from year to year (particularly in the upper aquifer), but no consistent trend developed towards either more strongly or more mildly reducing conditions. Occasional detections of sulfide, widespread detections of methane, and frequent detections of dissolved H2 at concentrations greater than 1 mg/L indicated that the strongly reducing conditions of sulfate-reduction and methanogenesis were present in much of the upper aquifer beneath the landfill and in parts of the intermediate aquifer downgradient from the landfill. These redox conditions are most favorable for reductive dechlorination of all chloroethenes (table 2). Mildly reducing conditions, which are moderately favorable for reductive dechlorination of TCE but are less favorable for reductive dechlorination of cis-DCE and VC, are present in the remainder of the contaminated parts of the upper and intermediate aquifers. Mildly reducing ground water is favorable for microbial oxidation of VC (and to a lesser extent, cis-DCE), so the lack of strongly reducing conditions and reductive dechlorination throughout the system likely would not lead to an accumulation of VC in downgradient wells.
CVOC data collected by the USGS from piezometers, selected wells, and passive-diffusion samplers at OU-1 in June 2004 are shown in table 3. Data from wells and piezometers are grouped and presented as upgradient, northern plantation, southern plantation, and intermediate aquifer. CVOC data from all passive-diffusion samplers are presented in a final group labeled “marsh (passive-diffusion samplers)” in table 3. Complete analytical results (including data qualifiers) are available from the USGS National Water Information System web site http://nwis.waterdata.usgs.gov/wa/nwis/qwdata. A cumulative summary of the 1991–2000 CVOC data collected by various agencies at OU-1 is available in Dinicola and others (2002). Cumulative summaries of the 2001–03 CVOC data collected by the Navy and the USGS are available in CH2M Hill Constructors, Inc. (2005) and Dinicola and Huffman (2004), respectively. Selected historical chloroethene data previously published are presented graphically in this report. In the following discussions of trends in contaminant concentrations, there is uncertainty that results from relatively few samples available at some sites, and from varying minimum reporting levels for selected contaminants. Trends in contaminant concentrations are uncertain because of the relatively few sample results available for some sites and the various minimum reporting levels for selected contaminants.
In the upper aquifer (including all upgradient, northern plantation, southern plantation, and marsh sites), the 2004 concentrations of total CVOCs (table 3 and fig. 5) generally were less than those measured in 1999–2000 (fig. 6). The term “total CVOCs” as used in this report is the sum of the six chloroethene and three chloroethane compounds of concern at OU-1. Lower concentrations were most apparent throughout the northern plantation, in the northern part of the southern plantation, and in areas outside the plantation boundaries. Lower concentrations in areas outside the plantation boundaries confirm that contaminant attenuation processes other than phytoremediation are active at the landfill. The 2004 total CVOC concentrations remained exceptionally high (49,000–103,000 µg/L) in the southern part of the southern plantation at sites MW1-4, P1-7, and P1-9. Non-aqueous phase liquids in the subsurface likely caused both the persistence and magnitude of CVOC concentrations in that area. USGS did not determine CVOC concentrations in the intermediate aquifer during 2004, but the Navy measured concentrations similar to those measured since 2000.
Chloroethenes were not detected at upgradient wells MW1-10 and MW1-33 since monitoring began at those sites in 1996 (table 3 and Dinicola and others, 2002; Dinicola and Huffman, 2004; and CH2M Hill Constructors, Inc., 2005). TCE was detected once during 1999 at upgradient well MW1-3 at a concentration of 0.70 µg/L, although no other chloroethenes have been detected there since monitoring began in 1991 (Dinicola and others, 2002; Dinicola and Huffman, 2004; and CH2M Hill Constructors, Inc., 2005).
Chloroethene concentrations at the three most contaminated sample sites in the northern plantation (1MW-1, MW1-2, and P1-4) generally decreased over the periods of record (fig. 7). Chloroethene concentrations at piezometers P1-1, P1-3, and P1-5 in the northern plantation have also decreased substantially since 1999 (table 3 and Dinicola and others, 2002; Dinicola and Huffman, 2004; and CH2M Hill Constructors, Inc., 2005).
Decreasing concentration ratios of more highly chlorinated compounds to less chlorinated compounds (TCE:cis-DCE and cis-DCE:VC) over time indicate that biodegradation is a substantial cause for the downward trend in contaminant concentrations beneath the northern plantation. The TCE:cis-DCE ratios for ground water beneath the northern plantation exceeded 1 at only two sites (P1-1 and P1-5) during only one year (Dinicola and others, 2002; Dinicola and Huffman, 2004; and CH2M Hill Constructors, Inc., 2005). At all other sites, TCE:cis-DCE ratios were less than 0.1 and have decreased over time, indicating continued reductive dechlorination of TCE. The cis-DCE:VC ratios, as well as cis-DCE concentrations, decreased over time at most northern plantation sites (fig. 7) indicating continued reductive dechlorination of cis-DCE. Although VC concentrations have not consistently decreased throughout the vicinity of the northern plantation, they also have not increased despite the continued production of VC through reductive dechlorination of cis-DCE. VC appears to be biodegraded at a rate similar to its rate of production. Reductive dechlorination of VC is reliably indicated by ethane plus ethene concentrations as high as 60 µg/L in the northern plantation (table 3). Microbial oxidation of VC (and cis-DCE to a lesser extent) also may be occurring in the iron- and manganese-reducing parts of the aquifer, but no unique diagnostic byproducts reliably indicate that process.
Chloroethenes migrating from beneath the northern plantation are further biodegraded beneath the marsh before discharging to surface water. Chloroethene concentrations decreased between 2000 and 2004 in six of the eight passive diffusion sampler sites near the northern plantation (marsh sites N-1 through N-5, N-7, and N-8; table 3), and the concentration at the site with the highest 2000 concentration (570 µg/L at N-3) decreased to about 2 µg/L in 2004 (table 3). The Navy measured VC concentrations in a seep 50 ft north of well MW1-2 that decreased from a high of 420 µg/L during 1996 to less than 0.5 µg/L during 2004 (CH2M Hill Constructors, Inc., 2005). Methane concentrations during 2004 at the passive-diffusion sampler sites were as high as 290 mg/L (table 3), suggesting the presence of methanogenic redox conditions in marsh ground water that are favorable for reductive chlorination of all chloroethenes. Ethane plus ethene concentrations as high as 68 µg/L at passive-diffusion sampler sites near the northern plantation are reliable evidence for reductive dechlorination of VC (table 3).
Chloroethene concentrations at all sampling sites in the northern part of the southern plantation (MW1-5, MW1-16, P1-6, P1-8, and P1-10) decreased by one to three orders of magnitude over the periods of record (fig. 8; data for MW1-5 not shown). In contrast, chloroethene concentrations at sites in the southern part of the southern plantation (MW1-4, P1-7, and P1-9) remained high with no consistent downward trend (fig. 9).
Dilution may cause some of the measured decrease in chloroethene concentrations in the northern part of the southern plantation. Coincidental downward trends in chloride (table 6) and chloroethene concentrations (fig. 8) were measured at sites MW1-16, P1-6, P1-8, and P1-10, although no downward trend in chloride concentrations developed at the upgradient site MW1-20. These data indicate that ground water recharge with relatively low chloride concentrations led to increased contaminant dilution in the northern part of the southern plantation. Recharge presumably is the result of pavement removal for tree planting in 1999. However, chloroethene concentrations decreased proportionally more compared to chloride concentrations, so the observed downward trends in chloroethene concentrations can be partly attributed to increased dilution and partly attributed to biodegradation. In contrast, chloride concentrations were relatively consistent over time in the southern part of the southern plantation at sites MW1-4, P1-7, and P1-9 (table 6), which does not indicate an increase in dilution over time.
Although TCE and other chloroethene concentrations decreased over time in ground water beneath the northern part of the southern plantation (fig. 8), no corresponding decreases in TCE:cis-DCE ratios have been measured to clearly indicate continued reductive dechlorination of TCE (calculated from data in Dinicola and others, 2002; Dinicola and Huffman, 2004; and CH2M Hill Constructors, Inc., 2005). Continued reductive dechlorination of cis-DCE was indicated by decreasing cis-DCE:VC ratios (fig. 8), and reductive dechlorination of VC is reliably indicated by ethane plus ethene concentrations as high as 217 µg/L (table 3). A likely explanation for those data is that reductive dechlorination of all chloroethenes continued, but that a continuing (and waning) source of dissolved TCE from residual non-aqueous phase contaminants was present. As was noted for the northern plantation, microbial oxidation of VC (and cis-DCE to a lesser extent) also may be occurring in the iron- and manganese-reducing portions of the aquifer, but there are no unique microbial-oxidation byproducts by which to diagnose that process. To reliably attribute the measured contaminant concentration trends to phytoremediation activities is difficult because of the substantial inter-annual variation in chloroethene concentrations measured at site MW1-16 (fig. 8), the only nearby site with pre-1999 data. The post-1999 variation at site MW1-16 in part may be due to more frequent than annual sampling and to different sampling techniques used by the Navy and the USGS.
In contrast to all other parts of the landfill, chloroethene concentrations in the southern part of the southern plantation have been exceptionally high, with no consistent decrease (fig. 9). No downward trends were revealed in TCE:cis DCE or cis-DCE:VC ratios at sites MW1-4, P1-7, and P1-9 to clearly indicate continuing reductive dechlorination of TCE, even though reductive dechlorination to non chlorinated end products was reliably indicated by ethane plus ethene concentrations ranging from 216 to 556 µg/L (table 3). A likely explanation for those data is that reductive dechlorination of all chloroethenes is ongoing, but that a continuing persistent source of dissolved TCE is present.
Beneath the marsh near the southern plantation, chloroethene concentrations increased between 2000 and 2004 at all three passive-diffusion sampler sites with data available for both years (marsh sites S-4 through S-6, table 3). Chloroethene concentrations in ground water discharging to the 100-foot long reach in the southern part of the marsh between sites S-4 and S-6 were at least 300 times greater than concentrations in ground water discharging elsewhere in the marsh (table 3). Total CVOC concentration in the most contaminated passive-diffusion sampler site (S 4) increased from 25,000 to 40,000 µg/L (Dinicola and others, 2002, and table 3). For comparison, the highest total CVOC concentrations measured beneath the landfill (P1-7) increased from 75,000 to 103,000 µg/L between 2000 and 2004 (Dinicola and others, 2002, and table 3), a proportionally similar amount. TCE:cis-DCE and cis-DCE:VC ratios between P1-7 in the landfill (fig. 9) and S-4 in the marsh (ratios not shown) decreased substantially during both 2000 and 2004, indicating substantial reductive dechlorination of both TCE and cis-DCE between the landfill and the marsh. Reductive dechlorination of VC beneath the marsh is indicated by the 2004 ethane plus ethene concentrations of 1,122 µg/L at S-4 in the marsh and 556 µg/L at P1-7 in the landfill (table 3). Methane concentrations at S-4 in the marsh during 2004 were 530 mg/L, indicating methanogenic redox conditions in marsh ground water that are highly favorable for reductive chlorination of chloroethenes.
Considering the southern plantation as a whole, the only clear and consistent trends are decreasing chloroethene concentrations measured at the northern part of the southern plantation. Both dilution and biodegradation are reducing concentrations in that vicinity, and there appears to be only a small and decreasing amount of non-aqueous phase chloroethene source, likely TCE, to continue replenishing ground-water contamination. Biodegradation also is active in the southern part of the southern plantation, but chloroethene concentrations did not decrease. A more substantial amount of non-aqueous phase chloroethene source appears to continue replenishing ground-water contamination in that vicinity, and no measurable increase in dilution followed the 1999 asphalt removal. Although chloroethene concentrations increased during 2004 at the marsh passive-diffusion sampler sites, the 2000 and 2004 data were not sufficient to define a trend in contaminant migration beyond the southern plantation. Chloroethene concentrations in marsh surface water were measured at least twice per year by the Navy at site MA-12 just upgradient from the marsh pond (fig. 10). Concentrations increased between 1996 and early 1999 (after pavement removal), decreased through late 2003, and began to increase again during 2004. More data are needed to determine if the 2004 concentrations at site MA-12 represent an upward trend in contaminant discharge to the marsh.
TCE concentrations measured in the intermediate aquifer near the downgradient margin of the landfill at well MW1-25 decreased between 1996 and 2004 (fig. 11). TCE concentrations also generally decreased at well MW1-28 with the exception of a measurement in 2002. Concentrations of cis-DCE and VC measured in those wells increased slightly between 1996 and 2000, but from 2001 to 2004, concentrations decreased or remained the same. Farther downgradient in the intermediate aquifer beneath the Highway 308 causeway (wells MW1-38 and MW1-39), TCE was never detected (table 3 and Dinicola and others, 2002; Dinicola and Huffman, 2004; and CH2M Hill Constructors, Inc., 2005). Concentrations less than 2 µg/L of cis-DCE and VC were consistently detected in the shallower of the adjacent intermediate aquifer wells (MW1-39) with no consistent trend in those concentrations (Dinicola and others, 2002; Dinicola and Huffman, 2004; and CH2M Hill Constructors, Inc., 2005).
Biodegradation of chloroethenes in the most contaminated part of the intermediate aquifer (represented by wells MW1-25 and MW1-28) was uncertain based on the 1995–2000 data (Dinicola and others, 2002). Data through 2004 reliably indicate that some reductive dechlorination in the intermediate aquifer occurred. In addition to decreases in TCE concentrations between 2000 and 2004 at wells MW1-25 and MW1-28 (fig. 11), TCE:cis-DCE ratios decreased from 0.02 to 0.005 at MW1-25 and from 0.001 to 0.0004 at MW1 28 during that period (calculated from data in table 3 and CH2M Hill Constructors, Inc., 2005). Those trends indicate reductive dechlorination of TCE. Reductive dechlorination of cis-DCE and VC is less certain. Concentrations of cis-DCE decreased moderately since 2002 after a more substantial increase measured between 1996 and 2000. In contrast, cis-DCE:VC ratios consistently increased during 1996–2004 (fig. 10) due to relatively stable VC concentrations. Ethane plus ethene concentrations ranging from 19.7 to 29.8 µg/L at MW1-25 and MW1-28 are reliable evidence for reductive dechlorination of VC (table 3), but those concentrations are low compared to many upper aquifer sites. Together, those data indicate some, but not substantial, biodegradation. Data from the contaminated well at Highway 308 (MW1-39) neither support nor refute ongoing biodegradation. VC concentrations during 1996–2004 ranged between 0.76 and 2 µg/L, and cis-DCE concentrations ranged between 0.32 and 0.56 µg/L (CH2M Hill Constructors, Inc., 2005).
The rates at which the chloroethene mass at OU-1 was degraded in ground water in the upper aquifer or was discharged to surface water were estimated using 2004 data and compared to estimates made using 1999–2000 data. Rates were calculated according to a flux-based approach that estimated the flux of chloroethenes across two parallel transects oriented perpendicular to ground-water flow direction in the upper aquifer (fig. 12). The upgradient (landfill) transect follows the western and southern margin of the landfill, and the downgradient (marsh) transect follows the creek and pond in the adjacent marsh. The difference in chloroethene fluxes estimated for the two transects is an estimate of the mass of chloroethenes degraded in ground water over a given duration. The mass of non-dissolved chloroethenes in the landfill (non-aqueous phase liquids) is unknown, so calculations considered only the dissolved contaminants mass. Not enough data are available to estimate chloroethene mass degradation in the intermediate aquifer or chloroethene discharge to surface water from the intermediate aquifer.
To better characterize chloroethene degradation and discharge in different parts of the study area, chloroethene fluxes were calculated for five pairs of parallel sub-transects along the landfill and marsh transects. The upgradient landfill transect is bounded by well MW1-18 to the north and well MW1-4 to the south, and is divided into sub-transects as indicated in table 4 and figure 12. The downgradient transect is bounded by the passive-diffusion sampler sites N-1 and S-6, and is divided into similar sub-transects as indicated in table 4 and figure 12. Chloroethene fluxes were calculated by multiplying estimated ground-water fluxes by measured chloroethene concentrations in wells or passive-diffusion samplers on the transects. Ground-water fluxes across each landfill sub-transect were estimated by URS Consultants Inc. (1997b) using measured or estimated transmissivity and hydraulic gradient data. It was assumed that the ground water fluxes across corresponding marsh sub-transects were the same as those estimated for the landfill sub-transects, and that all ground water discharged to marsh surface water immediately after passing the marsh transect. Measured chloroethene concentrations at sample sites along each sub transect were proportionally weighted according to distance represented by each site to calculate average concentrations for each sub-transect.
The chloroethene mass degraded in ground water per day was calculated as the difference between fluxes across the landfill transect and across the marsh transect. Calculations for DCE (including both cis- and trans-isomers) assumed that the mass of degraded TCE resulted in the formation of a molar equivalent amount of DCE through reductive dechlorination, and the calculations for VC assumed that the mass of degraded DCE resulted in the formation of a molar equivalent amount of VC through reductive dechlorination. The amount of PCE available in the landfill to degrade into TCE was considered negligible. The rate of chloroethene discharge to surface water was conservatively assumed to be the chloroethene flux calculated for the marsh transect. Calculations were done using measured chloroethene data collected in 1999–2000 and 2004. The 1999–2000 results presented in this report differ somewhat from those reported by Dinicola and others (2002) because the average chloroethene concentrations for sub transects were calculated for a different set of wells.
Flux calculations required the following assumptions:
· Ground-water sampling sites represented conditions throughout the study area—This assumption is reasonable for the relatively dense spatial network of wells, piezometers, and passive-diffusion samplers in the study area, although the vertical distribution of contaminants was not as well represented; uniform contaminant concentrations throughout the saturated thickness of the upper aquifer were assumed.
· Ground-water flow, the supply of dissolved contaminants, and contaminant degradation rates were all at steady-state conditions—Steady-state ground-water flow is a reasonable assumption based on water-level data measured since 1996, and post-1999 water-level data, which do not indicate that phytoremediation has affected the flow and ground-water gradients (URS Greiner, Inc., 2000; CH2M Hill Constructors, Inc., 2004). A relatively steady supply of dissolved contaminants is indicated by stable or slowly decreasing CVOC concentrations in the most contaminated landfill wells. Steady-state contaminant degradation rates are suggested by generally consistent redox conditions between 1996 and 2004.
· No substantial contaminant loss in ground water was caused by sorption, volatilization, plant uptake beneath the phytoremediation plantations, or advective transport to the intermediate aquifer—Volatilization losses probably were minimal because volatilization of chemicals from ground water is greatly constrained by the rate of vapor transport upward through the unsaturated zone. Initially, sorption losses may have been significant in the organic-rich marsh sediments, but contaminants flowed through those sediments for many decades so the bulk of the sorption capacity probably was filled long ago. Contaminant uptake by plants from the unsaturated zone was detected beneath the plantations, but no evidence was detected indicating that contaminants are drawn out of the saturated zone (URS Greiner, Inc., 2002). Most of the advective transport of contaminants from the upper to the intermediate aquifer likely occurs upgradient from the landfill transect, so that advective transport will not affect the mass degradation calculations. The only downward gradient for ground-water flow between the upper and intermediate aquifers is beneath the northeastern one-third of the landfill, and an upward gradient is beneath the remainder of the landfill (URS Consultants, Inc., 1997a). Any minor contaminant losses through these mechanisms are implicitly included in the calculated mass degradation rates.
Flux estimates based on 2004 data indicate that most dissolved-phase chloroethene mass in the upper aquifer beneath the landfill was degraded before it discharged to surface water (table 4). Of the 72 g/d of chloroethene flux measured at the landfill transect during 2004, only 13 percent (9.3 g/d) migrated and discharged to surface water in the marsh; therefore, the total chloroethene flux was reduced by 87 percent due to biodegradation in the upper aquifer. Flux estimates based on 1999–2000 data indicated that a larger dissolved contaminant flux from the landfill (94 g/d) was reduced by a larger percentage (93 percent) in the upper aquifer and resulted in less chloroethene discharged to surface water (6.5 g/d) during 1999–2000.
During 1999–2000 and 2004, the chloroethene with the greatest calculated flux into marsh surface water was cis-DCE, followed by VC and TCE. However, the mass degradation rates for cis-DCE (54 and 72 g/d) also were the greatest, followed by VC and TCE. The rate of DCE degradation was about 50 percent greater than the flux of DCE across the landfill transect, but DCE still reached the marsh because it was continually created by reductive dechlorination of TCE. The rate of VC degradation was about 10 times greater than the flux of VC across the landfill transect, but VC still reached the marsh because it was continually created by reductive dechlorination of DCE. During both periods, nearly all TCE flux to the marsh also was from the southern plantation. Most DCE and VC flux to the marsh also was from the southern plantation, although measurable fluxes also came from the northern plantation.
Although the mass degradation rates were greater during 1999–2000 and the rates of discharge to surface water were greater during 2004, it cannot be concluded that biodegradation has become less effective. Flux estimates are extremely sensitive to measured chloroethene concentrations at the few highly contaminated wells and passive-diffusion samplers in and near the southern plantation, and the interannual variability in measured concentrations at those sites was high. Estimated mass degradation rates were exceptionally high during 1999–2000, in large part due to the large mass of degradable chloroethenes. At P1-9, for example, total CVOC concentrations were 160,000 µg/L during June 2000 and 75,000 µg/L during June 2004 (figs. 5 and 6). Similarly, calculated rates for chloroethene discharge to surface water were exceptionally high during 2004 in large part because of exceptionally high chloroethene concentrations measured at site S-4, where total CVOC concentrations were 25,000 µg/L during June 2000 and 40,000 µg/L during June 2004 (figs. 5 and 6). In contrast to the site S-4 data, total chloroethene concentrations measured in surface water at MA 12 were 920 µg/L in June 2000, compared to 560 and 830 µg/L in April and October 2004 (CH2M Hill Constructors, Inc., 2005). Too few data are available from site S-4 to permit characterization of the interannual variability in chloroethene concentrations.
The accuracy of estimated mass degradation rates was indirectly assessed by evaluating the accuracy of estimated chloroethene flux to surface water in the southern part of the marsh. The accuracy of estimated chloroethene fluxes to surface water presented in table 4 was evaluated by comparing those estimates to an independent estimate of chloroethene flux to surface water in the southern part of the marsh. The independent estimate was determined by multiplying measured chloroethene concentrations in surface water at site MA-12 by estimated surface-water discharge at site MA-12 (table 5). No surface-water discharged into the marsh from surrounding areas during non-storm periods, so surface-water flow in the marsh was entirely due to ground-water discharge. Therefore, surface water flowing past site MA-12 included 0.2 gal/min of contaminated ground water from the landfill combined with 1.1 gal/min of uncontaminated ground water that discharged to the marsh from areas south of the marsh creek (URS Consultants Inc., 1997b).
The TCE flux to marsh surface water shown in table 4 was substantially underestimated for 1999–2000 and 2004, indicating that the TCE mass degradation rates in table 4 likely were overestimated. Clearly, not all TCE flux from the landfill was represented by the measured concentrations from the passive-diffusion sampler sites. The measured TCE concentration in surface water at site MA-12 in June 2000 was 110 µg/L, while the highest TCE concentration measured in passive-diffusion samplers during 2000 was only 49 µg/L (Dinicola and others, 2002). The highest probable TCE concentration in passive-diffusion samplers during 2004 was poorly quantified [reported at less than 1,000 µg/L for S-4 (table 3)], so a similar comparison could not be made.
In contrast to TCE, the flux of VC to marsh surface water shown in table 4 was overestimated by nearly 600 percent for 2004, indicating that the VC mass degradation rates in table 4 likely were underestimated. The measured VC concentrations in surface water at site MA-12 during 2000 and 2004 ranged from 46 to 250 µg/L (fig. 10), although the highest VC concentrations measured in passive-diffusion samplers ranged from 5,600 to 17,000 µg/L (Dinicola and others, 2002 and table 3). There appears to be additional attenuation of VC concentrations between shallow ground water in the marsh and site MA 12. The likely mechanisms behind that attenuation include microbial degradation of VC in the upper 2 ft of organic-rich marsh sediments, and volatilization of VC from surface water before it reaches site MA-12.
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