Environ. Sci. Technol. 2007, 41, 1384-1389
Biologically Enhanced Mass Transfer of Tetrachloroethene from DNAPL in Source Zones: Experimental Evaluation and Influence of Pool Morphology KENT C. GLOVER,* JUNKO MUNAKATA-MARR, AND TISSA H. ILLANGASEKARE Environmental Science and Engineering Division, Colorado School of Mines, Golden, Colorado 80401
High-saturation pools of dense nonaqueous phase liquid (DNAPL) are long-term sources of groundwater contamination at many hazardous-waste sites. DNAPL pools consist of a high saturation zone with slow dissolution overlaid by a transition zone with lower saturations and more rapid dissolution. Effects of biological activity on pool dissolution must be understood to evaluate and implement bioremediation strategies. Bioenhanced dissolution of tetrachloroethene (PCE) in transition zones of high-saturation pools was investigated in a custom-designed 5-cm flow cell. Experiments were conducted to characterize mass transfer following DNAPL emplacement, with and without an active microbial culture capable of reductive dehalogenation. For average pool saturations e0.55, mass transfer during biodegradation was enhanced by factors of 4-13, due primarily to high mass flux of PCE degradation products. However, at an average pool saturation of 0.74, mass transfer was enhanced by factors less than 1.5. Mass transfer was significantly greater from pools with an observable transition zone than without. Advective flow through multiphase transition zones enhanced dissolution and biological activity. These laboratory-scale experimental results suggest that biotechnologies may be effective remediation strategies for depletion of source zones within pool transition zones.
Introduction Use of chlorinated solvents has resulted in extensive groundwater contamination. Many chlorinated solvents are denser than water and migrate by gravity through the saturated zone. During migration, pore-scale capillary forces retain a portion of dense nonaqueous phase liquid (DNAPL) as discontinuous ganglia at residual saturations (1, 2). Large DNAPL volumes also are retained as pools (i.e., zones with DNAPL at greater than residual saturation). Pooling results from DNAPL accumulation on low-permeability layers (3) or capillary entrapment within coarse-grained lenses (4, 5). Spill experiments in large tanks filled with heterogeneous porous media (6, 7) show that pools form thin layers that are virtually impossible to remove by pumping when density differences between DNAPL and water are large, as with tetrachloro* Corresponding author phone: (303) 273-3421; fax (303) 2733411; e-mail:
[email protected]. 1384
9
ENVIRONMENTAL SCIENCE & TECHNOLOGY / VOL. 41, NO. 4, 2007
ethene (PCE). Spatially averaged, thin pools can appear as a source zone at residual saturation. Field-scale detection of thin pools is problematic. However, groundwater flow and DNAPL dissolution in the vicinity of thin pools is distinct from residual zones (5, 8). Long-term persistence of DNAPL is closely tied to dissolution from thin pools (8). This paper focuses on bioenhanced dissolution from pools of this type. DNAPL pools often are conceptualized as homogeneous zones of high saturation with sharp boundaries (9-12). However, laboratory, field, and theoretical research (3, 4, 13, 14) indicate pools are bounded by transition zones where saturation changes from high to residual values. Detailed mapping shows that saturation increases and pore-water velocity decreases with depth into a pool (15). In PCE DNAPL pools with average saturation of 0.30-0.70, transition zones are millimeters to centimeters deep. Entrapped DNAPL mass dissolves slowly into flowing groundwater and is a long-term contamination source, posing significant remediation challenges (2, 16). For heterogeneous sources, long-term persistence of contaminants is controlled by slow dissolution of DNAPL from pools (8, 17-19). Slow dissolution results from decreased relative permeability which causes groundwater to flow predominantly around rather than through pools (5, 20). As a result, dissolution occurs at pool edges (9, 20). Bioremediation of chlorinated DNAPL occurs in the aqueous phase by anaerobic microbes that use various substrates as electron donors and dissolved chlorinated organics as terminal electron acceptors for microbial respiration (21). Laboratory studies have demonstrated the ability of dehalorespiring microbes to metabolize high concentrations of PCE (22, 23) and to enhance dissolution of PCE from DNAPL at residual saturation (24-29). Application of enhanced bioremediation to chlorinated DNAPL zones at field scale also has been reported (30, 31). Dissolution can be enhanced by biodegradation acting as a reaction sink near DNAPL-water interfaces to increase PCE dissolution gradients. PCE degradation products trichloroethene (TCE), cis-1,2-dichloroethene (DCE), vinyl chloride (VC), and ethene have much greater solubility than PCE. Therefore, bioactivity can result in decreased aqueous concentration of PCE near DNAPL-water interfaces and increased concentration of less hydrophobic metabolites that are transported from the source zone. Dissolution of PCE also can be enhanced by biosurfactants or electron-donor co-solvency, with subsequent aqueous-phase biodegradation. Regardless of dissolution mechanism, overall DNAPL depletion flux is greater than would be observed without bioactivity (32). Effects of biodegradation on PCE dissolution from pools are poorly understood. If bioactivity significantly enhances dissolution and decreases source longevity, while simultaneously degrading PCE to less toxic compounds, bioremediation could be highly effective for source-zone treatment. Effects of bioactivity on dissolution are likely to depend on saturation distribution within pools. If a pool forms as a homogeneous zone of high saturation with sharp boundaries, dissolution may be enhanced to a limited degree. Bioenhancement would be limited by electron-donor transport to pool boundaries and the limited interfacial area available for PCE dissolution. Abiotic dissolution from pools with sharp boundaries is controlled by transverse hydrodynamic dispersion (10, 33) and a similar control for bioenhanced dissolution might be expected. Recent bioenhanced dissolution experiments with LNAPL pools of dodecane and toluene showed that mass transfer was enhanced only by a 10.1021/es060922n CCC: $37.00
2007 American Chemical Society Published on Web 01/12/2007
TABLE 1. Characteristics and Conditions of DNAPL Dissolution Experiments lower/upper average DNAPL number sand volumea saturation (Sn) test name of tests (mL) porosity in lower sand A1 A2
4 5
4.5/12.3 3.9/12.9
0.37 0.40
A3 A4 A5 A6 A7
3 3 3 3 3
4.0/12.8 3.4/11.4 3.8/11.8 3.8/11.4 3.2/12.2
0.43 0.42 0.42 0.38 0.40
B1 B2
4 5
4.5/12.3 3.9/12.9
0.37 0.40
B3
3
4.0/12.8
0.43
a
flow rates (mL/min)
comment
Abiotic Dissolution Experiments 0.55 0.012, 0.008, 0.004, 0.012 inactive biomass 0.25 0.022, 0.014, 0.017, 0.014, 0.005 inactive biomass, methanol varied, X-ray scan 0.74 0.012, 0.009, 0.0034 inactive biomass, X-ray scan 0.67 0.084, 0.168, 0.253 no biomass 0.88 0.084, 0.168, 0.253 no biomass 0.45 0.084, 0.168, 0.253 no biomass, dye tests, X-ray scan 0.30 0.084, 0.168, 0.253 no biomass Bioenhanced Dissolution Experiments 0.55 0.012, 0.008, 0.004, 0.012 inflow methanol, 6.25 mmol/L 0.25 0.007, 0.013, 0.025, 0.008, 0.005 methanol given in supplemental Table S2, X-ray scan 0.74 0.013, 0.008, 0.0038 methanol 6.25 mmol/L; X-ray scan
Excludes volume of inflow and outflow capillaries.
factor of 2 (34). The LNAPL pools were constructed at very high saturation, resulting in sharp pool boundaries. In a pool with a transition zone overlying a high saturation zone, very different behavior might be expected. DNAPL saturation in the transition zone increases from residual saturation at the top of the pool toward higher saturations at greater depths. Relative permeability with respect to water is larger in the transition zone than in deeper parts of the pool. Where pool entrapment occurs in coarse-grained lenses, transition zone hydraulic conductivity may be larger than that of adjacent fine-grained material. As a result, electron donor may preferentially migrate within the transition zone. Combined with larger DNAPL-water interfacial area, biodegradation within transition zones may significantly enhance dissolution. The research objectives of this study were to experimentally determine the effects of biodegradation on mass transfer from DNAPL pools containing PCE and to identify the significance of pool transition zones to bioenhanced dissolution. Flow-cell experiments with a range of pool saturations and flow rates were conducted with and without active biodegradation. Spatial variability in DNAPL saturation was quantified by X-ray attenuation techniques. Comparing saturation-distribution data with dissolution-experiment results, the role of transition zones in bioenhanced dissolution was clarified. The experiments were part of a larger effort to provide quantitative tools for modeling bioenhanced dissolution. This larger effort included quantitative modeling of flow-cell experiments, and experiments and modeling of large tanks to develop upscaling strategies.
Experimental Section Additional details for experimental methods are provided as Supporting Information. Culture Development. The KB-1 mixed culture of dechlorinators (35) was used in experiments. Prior to use, batch cultures were adapted to PCE at high aqueous concentration (0.72 mmol/L) and were maintained in anaerobic nutrient medium using established procedures (36). Flow Cell Configuration. The spheroidal flow cell was 5 cm long and 2.5 cm in diameter with access ports on horizontal and vertical axes. A capillary trap for PCE-DNAPL was created by filling the lower part of the 16.5 mL flow cell with coarse-grained sand and the overlying volume with medium- to fine-grained sand. Laboratory tests show that water-saturated hydraulic conductivity (Ks) of the upper sand was 2.8 cm/min while Ks of the lower sand was 42 cm/min (4, 7). Both sands were sterile and free of organic matter
prior to packing. The flow cell was inoculated with KB-1 to create a homogeneous biomass distribution. Nutrient medium, methanol, and PCE were then injected through the flow cell and effluent was analyzed for PCE and degradation products. When PCE degradation approached steady conditions, a DNAPL pool was created within the lower portion of the flow cell. The use of PCE DNAPL is a simplification from multicomponent DNAPL often found at field sites. Anaerobic conditions were maintained throughout each experiment. X-ray Attenuation System. Spatial distribution of DNAPL was monitored with a high-resolution X-ray attenuation system. Differential path lengths of sand, water, and DNAPL were determined at 2 mm vertical and 25 mm horizontal spacing within the flow cell. Material path lengths are measures of porosity and fluid saturation. Analysis of photon attenuation spectra was based on an application of the BeerLambert law for multiple materials and an equivalent number of lumped-energy ranges (6, 37). The resulting matrix expression was solved for path lengths given a set of attenuation coefficients determined from calibration samples of known path length. Experimental Design. Three sets of paired biotic-abiotic experiments were conducted (Table 1: B1-A1, B2-A2, and B3-A3). For each set, bioenhanced dissolution experiments were conducted at 3-5 flow rates, as were an equal number of abiotic tests at comparable flow rates. Average DNAPL saturation within the lower sand was established as 0.25, 0.55, and 0.74 for the three sets, respectively. Inflow water for all experiments was nutrient medium with specified methanol concentration added as electron donor. In the first and third set of experiments, a uniform methanol concentration was used (6.25 mmol/L). This concentration exceeded biodegradation needs of batch microcosms operated with PCE concentrations of 0.72 mmol/ L. In the second set, methanol concentration was varied between experiments from 2.6 to 9.4mmol/L. Flow rates provided seepage velocities consistent with field conditions (38, 39). Each set of experiments consisted of (1) preparing and inoculating a flow cell, (2) injecting a known amount of PCE to form a DNAPL pool within the lower sand, (3) conducting bioenhanced dissolution experiments at selected flow rates and methanol concentrations, (4) introducing oxygen to the flow system to stop dechlorination, (5) conducting abiotic dissolution experiments with inactive biomass at flow rates used in the bioenhanced experiments, and (6) characterizing biomass by destructive sampling. X-ray attenuation data were VOL. 41, NO. 4, 2007 / ENVIRONMENTAL SCIENCE & TECHNOLOGY
9
1385
FIGURE 1. Initial DNAPL saturation, averaged in transverse direction, in flow cell for test series A6 (average pool saturation ) 0.45). obtained periodically to map saturation profiles through the DNAPL pool. After inoculation, and again after DNAPL injection, each flow cell was operated with nutrientmethanol feed and effluent was monitored to ensure that KB-1 had acclimated to experimental conditions. Dissolution experiments began after acclimation. To compare results, flow rates used in abiotic tests were intended to be the same as those used in biotic tests. However, actual flow rates (Table 1) varied from this design when particulates in nutrient medium became suspended and lodged within inflow tubing. An additional 23 abiotic dissolution experiments were conducted to characterize transient response of the flow cell (Table 1, A4-A7). Sampling and Analysis Procedures. Water samples of flow-cell effluent were obtained until cumulative flow exceeded 20 pore volumes. Samples were analyzed by gas chromatograph. Chlorinated ethene concentrations approached quasi-steady-state conditions after 6-10 pore volumes. Results of DNAPL dissolution experiments indicated that PCE at high concentration inhibited biodegradation of cis-DCE within the flow cell. Samples contained negligible VC or ethene. Dye Tests. Dye-tracer tests were conducted during some abiotic dissolution experiments, prior to and following injection of DNAPL, to evaluate differences in water velocity in the upper sand, pool transition zone, and deeper portions of the DNAPL pool. In each test, non-adsorbing red food dye (#40) dissolved in water was added to the inflow stream and migration of the dye through the flow cell was recorded photographically.
Results DNAPL Pool Configuration. X-ray data showed a heterogeneous distribution of DNAPL saturation in each experiment (Figure 1). Layered transition zones were observed across pool interfaces for average saturations of 0.25 and 0.45. At average saturation of 0.74, a sharp interface was observed. For each flow cell, total DNAPL volume obtained by X-ray attenuation methods matched DNAPL volume measured during injection. Changes in DNAPL saturation during any experiment were less than 2%. In all but one set of experiments, all DNAPL was entrapped within the lower sand. Experiments B3/A3 were designed to evaluate pool dissolution at very high saturation. X-ray data showed a sharp pool boundary at the top of the lower sand. However, a small region of residual saturation was mapped near the outflow port. During dissolution tests, effluent water chemistry reflected this residual DNAPL and, to a degree, compromised assessments of dissolution from pools at very high saturation. Discussion of these limitations is provided in the Supporting Information. 1386
9
ENVIRONMENTAL SCIENCE & TECHNOLOGY / VOL. 41, NO. 4, 2007
FIGURE 2. Vertical profiles of DNAPL saturation in flow cell measured by X-ray attenuation and simulated by Brook-Corey relationship. Saturation profiles for each flow cell were determined by averaging X-ray data horizontally along the central portion of the flow cell (5-6 data points per horizontal plane). Resulting saturation profiles were fit to the Brooks-Corey retention function (40)
(Se)1/λ )
Pd Pc
where Pd is displacement pressure, λ is pore-size distribution index, Pc is capillary pressure, Se is effective water saturation equal to (Sw - Sr)/(1 - Sr), Sw is water saturation, and Sr is residual water saturation. The Brooks-Corey function is used widely to relate DNAPL and water content to capillary pressure. Because capillary pressure varies with depth in a pool, the function can be used to characterize vertical saturation profiles. The parameters Pd, λ, and Sr are properties of the sand that were estimated by regression analysis. Additional description of procedures used to model saturation profiles is provided in the Supporting Information. The resulting profiles are shown in Figure 2. Dye Tests. Dye tests conducted prior to DNAPL injection showed rapid transport in the lower coarse-grained sand and correspondingly short residence times; residence times in the upper sand were longer. Dye tests conducted after DNAPL injection showed rapid transport in the upper portion of the lower sand where DNAPL approached residual saturation, moderate rates of transport in the upper sand with no DNAPL, and slow or no transport in deeper portions of the lower sand where DNAPL saturation was high. Velocities within the flow cell were estimated with a purely advective transport model based on RT3D (41). Hydraulic properties for model development were obtained from laboratory tests (42) adjusted for saturation profiles (Figure 2). Though dye tests did not provide quantitative measures of tracer breakthrough, visible leading edges of tracer clouds provided a qualitative check on model results. Simulated velocity profiles were nearly uniform throughout the upper sand, increased significantly through the transition zone, and decreased as saturation increased with depth in the pool. Description of modeling is provided in the Supporting Information. Abiotic Dissolution. PCE was the predominant ethene chemical detected in effluent, indicating that dechlorination was not occurring. Abiotic tests conducted after tests with biological activity occasionally showed trace concentrations of degradation products TCE and DCE. In one test (A1), TCE
FIGURE 3. Quasi-steady-state effluent concentration of chlorinated ethene chemicals in flow cell experiments with biological activity under different average DNAPL saturation (Sn) conditions. Data labels for series B2 are methanol concentration (mmol/L). and DCE were a significant portion of total mass flux. This test was conducted immediately after introducing oxygenrich water to the flow cell and results were not used in subsequent analysis. PCE concentration and mass flux for abiotic experiments showed complex response to variations in DNAPL saturation and flow rate. Throughout this paper, mass flux was calculated based on effluent concentration, flow rate, and cross-sectional area of the flow cell. Detailed discussion of abiotic results is provided in the Supporting Information. Bioenhanced Dissolution. Biodegradation kinetics exerted a strong control on DNAPL dissolution from pools (Figure 3). The majority of chlorinated ethene mass in effluent was in the form of cis-DCE. Total chlorinated ethene concentration exceeded the solubility of PCE (approximately 1 mmol/L) in many experiments. This emphasizes the importance of high DCE solubility, compared to PCE solubility, in promoting bioenhanced dissolution. Data from B2 experiments showed that inflow methanol flux had a pronounced effect on DNAPL mass transfer (Figure 4). Experiments with inflow methanol flux greater than 10-5 mmol/min/cm2 showed no electron-donor limitation. When methanol flux dropped below 10-5 mmol/min/cm2, the proportion of degradation products TCE and cis-DCE within the total chlorinated ethenes declined, indicating a decrease in dechlorination (Figure 4b). As dechlorination decreased, so did the total mass flux of chlorinated ethenes (Figure 4a), indicating that bioenhanced dissolution was electron-donor limited. Mass transfer rates in bioenhanced dissolution experiments where methanol flux was greater than 10-5 mmol/ min/cm2 were greater than those observed in subsequent abiotic experiments (Figure 5). Mass transfer in experiments with average DNAPL saturation in the lower sand of 0.55 or less was enhanced by factors of 4-13. Greater enhancement occurred in experiments with slower inflow rates. At an average pool saturation of 0.74, mass transfer was enhanced by a factor less than 1.5. At high pool saturation, the thickness of the saturation transition zone was negligible, as was flow of water through the pool.
Discussion Experimental results demonstrate that dehalogenator activity can significantly enhance PCE depletion from a DNAPL pool. Consortia with reductive dechlorinators can grow at PCE concentrations approaching aqueous solubility and can populate source zones at residual saturation (24-29). Therefore, it is reasonable to expect dechlorinators to
FIGURE 4. Relationships of (a) chlorinated ethene mass flux and (b) daughter product - total dissolution flux ratio to methanol mass flux.
FIGURE 5. Steady-state mass flux of chlorinated ethene chemicals in flow cell effluent for dissolution experiments with and without biological activity; flux is normalized based on flow cell crosssectional area. populate pool transition zones. The degree to which pool depletion is enhanced by bioactivity reflects microbial activity within transition zones. Pools where dehalogenators are active within a transition zone would be expected to undergo depletion more rapidly than pools with minimal transition zones. The effective depth within a DNAPL pool to which a microbial culture will actively degrade PCE is likely to depend VOL. 41, NO. 4, 2007 / ENVIRONMENTAL SCIENCE & TECHNOLOGY
9
1387
FIGURE 6. Relationship of bioenhancement ratio to inflow rate and transition-zone thickness. on the supply of electron donor and nutrients. Because electron donor and nutrients are introduced to the subsurface at locations upstream of a DNAPL pool, mixing with electron acceptor (PCE) depends on groundwater velocity and dispersion near pool boundaries, particularly within the transition zone. Therefore, the effective depth of bioactive transition zones is likely to be determined, in large part, by the vertical profile of seepage velocity within the pool. Direct measurement of seepage velocity at the scale of a pool transition zone is not practical. However, the vertical profile of DNAPL saturation can be used as a basis for modeling the hydraulic conductivity profile in a pool. The Brooks-Corey relationship between relative permeability and effective water saturation (39) is widely used:
kr ) (Se)(2 + 3λ)/λ where kr is relative permeability. Hydraulic conductivity is calculated as the product of kr and Ks. Because Ks of the upper sand is an order of magnitude less than Ks of the lower sand, electron donor and nutrients introduced with inflow water preferentially migrated within the pool transition zone to a depth where water saturation decreased and hydraulic conductivity was less than Ks of the upper sand. Allowing for uncertainty in estimates of Ks and λ, an effective depth for preferential migration of electron donor and nutrients is assumed to be where kr < 0.05, estimated to occur at DNAPL saturation of 0.35 in this study. The effective thickness of bioactive transition zones is greatest for pools with low average DNAPL saturation, ranging in this study from 0.75 mm at the lowest saturation to negligible thickness at the highest. To assess the effects of transition zones on bioenhanced dissolution, a simple metric called the bioenhancement ratio was used. It is the ratio of steady-state molar concentration of PCE and degradation products observed in a biotic flowcell experiment to the steady-state concentration of PCE observed in an abiotic experiment conducted at the same conditions of flow rate and DNAPL saturation. Bioactivity had a greater effect on pools with an observable transition zone than in pools with a sharp saturation boundary (Figure 6). At lower flow rates, bioactivity enhanced mass transfer by approximately an order of magnitude. As flow rates increased, the bioenhancement ratio declined but remained greater than 3.5. In contrast, bioactivity had no significant effect on mass transfer from source zones at very high saturation where pools formed with sharp boundaries. Bioenhancement from residual zones has been predicted to increase with source-zone length (43) and a similar effect might be expected from pools. With increasing length, residence time within a source increases. By conducting experiments at a range of flow rates, the results in this paper 1388
9
ENVIRONMENTAL SCIENCE & TECHNOLOGY / VOL. 41, NO. 4, 2007
account for variations in residence time. However, increasing source length also increases interfacial surface area available for DNAPL dissolution and biodegradation. This effect is not included directly in the experiments. The primary focus of the experiments was to investigate relationships between pool transition zones and bioenhanced dissolution. A complete understanding of how pool morphology affects bioenhancement will require further experimentation with a range of source lengths, as well as consideration of complex heterogeneous sources involving multiple pools and residual zones at a range of scales. Experimentation and modeling along these lines has been completed and will be reported in subsequent papers. Nevertheless, results shown in Figure 6 suggest the following conceptual model to describe bioenhanced dissolution from pools. At low to moderate pool saturations, transition zones form with DNAPL saturation near residual values and hydraulic conductivity greater than the overlying sand. Electron donor may be transported rapidly by advection through these zones. Availability of PCE for biodegradation may be controlled solely by dissolution from the DNAPL phase, but transport of PCE into adjacent portions of the flow system is not required. Mixing of electron donor and PCE within transition zones promotes bioactivity, increasing PCE dissolution. Residence time in transition zones increases as flow rate decreases, allowing greater biodegradation. In pools with minimal transition zones, electron donor may be transported primarily around pools and mixing with PCE likely is controlled by transverse dispersion away from pool boundaries. Consequently, biodegradation at rates sufficient to depress aqueous-phase PCE concentration and increase DNAPL dissolution gradients likely occurs over a much smaller aquifer volume than in pools with observable transition zones. The traditional view of pools as aquifer zones with minimal advective transport suggests that biotechnologies would be ineffective for mass-transfer enhancement and source depletion. However, flow-cell experiments demonstrate that microbe-mediated reductive dechlorination can enhance dissolution following pool emplacement as much as an order of magnitude. Specifically, dissolution is enhanced in pools with observable saturation transition zones where hydrodynamics favor effective biodegradation and increased dissolution gradients. Long-term effectiveness of biotechnologies was not addressed by this research. However, these results help explain the effective treatment that is being seen at field scales with biotechnologies, although details of the processes have not been directly observed at field scales.
Acknowledgments This work was supported by the U.S. Department of Defense, Strategic Environmental Research and Development Program (CU-1294). We gratefully acknowledge provision of KB-1 by Dr. Elizabeth Edwards, University of Toronto.
Supporting Information Available Additional details including experimental methods for KB-1 development, flow-cell configuration, X-ray attenuation monitoring, sampling and analysis procedures, and biodegradation tests without DNAPL; tables of biodegradation test conditions, results of abiotic and bioenhanced dissolution tests, and comparison of biodegradation tests with other cultures. This material is available free of charge via the Internet at http://pubs.acs.org.
Literature Cited (1) Mercer, J. W.; Cohen, R. M. A review of immiscible fluids in the subsurface: properties, models, characterization and remediation. J. Contam. Hydrol. 1990, 6, 107-163.
(2) Schwille, F. Dense Chlorinated Solvents in Porous and Fractured Media; Lewis Publishers: Boco Raton, FL, 1988. (3) Kueper, B. H.; Abbot, W.; Fraquhar, G. Experimental observations of multiphase flow in heterogeneous porous media. J. Contam. Hydrol. 1989, 5, 83-95. (4) McWhorter, D. B.; Kueper, B. H. Mechanics and mathematics of the movement of dense non-aqueous phase liquids (DNAPLs) in porous media. In Dense Chlorinated Solvents and Other DNAPLS in Groundwater: History, Behavior and Remediation; Pankow, J. F., Cherry, J. A., Eds.; Waterloo Press: Portland, OR, 1996. (5) Saba, T.; Illangasekare, T. H. Effect of groundwater flow dimensionality on mass transfer from entrapped nonaqueous phase liquid contaminants. Water Resour. Res. 2000, 36, 971979. (6) Illangasekare, T. H.; Ramsey, J. L.; Jensen, K. H.; Butts, M. B. Experimental study of movement and distribution of dense organic contaminants in heterogeneous aquifers. J. Contam. Hydrol. 1995, 20, 1-25. (7) Illangasekare, T. H.; Yates, D. N.; Armbruster, E. J. Effect of heterogeneity on transport and entrapment of nonaqueous phase waste products in aquifers: an experimental study. J. Environ. Eng. 1995, 121, 572-579. (8) Parker, J. C.; Park, E. Modeling field-scale dense nonaqueous phase liquid dissolution kinetics in heterogeneous aquifers. Water Resour. Res. 2004, 40, W05109; doi: 10.1029/2003WR002807. (9) Jin, M.; Butler, G. W.; Jackson, R. E.; Mariner, P. E.; Pickens, J. F.; Pope, G. A.; Brown, C. L.; McKinney, D. C. Sensitivity model and design protocol for partitioning tracer tests in alluvial aquifers. Ground Water 1997, 36 (6), 964-972. (10) Sale, T. C.; McWhorter, D. B. Steady state mass transfer from single-component dense nonaqueous phase liquids in uniform flow fields. Water Resour. Res. 2001, 37, 393-404. (11) Roy, J. W.; Smith, J. E.; Gillham, R. W. Natural remobilization of multicomponent DNAPL pools due to dissolution. J. Contam. Hydrol. 2002, 59, 163-186. (12) Imhoff, P. T.; Pirestani, K. Influence of mass transfer resistance on detection of nonaqueous phase liquids with partitioning tracer test. Adv. Water Resour. 2004; doi: 10.1016/j.advwaters.2004.02.010. (13) Gerhard, J. I.; Kueper, B. H. Capillary pressure characteristics necessary for simulating DNAPL infiltration, redistribution and immobilization in saturated porous media. Water Resour. Res. 2003, 39, 1212; doi: 10.1029/2002WR001270. (14) Kueper, B. H.; Redman, D.; Starr, R. C.; Reitsma, S.; Mah, M. A field experiment to study the behavior of tetracholoroethylene below the water table: spatial distribution of residual and pooled DNAPL. Ground Water 1993, 31 (5), 756-766. (15) Moreno-Barbero, E.; Illangasekare, T. H. Influence of dense nonaqueous phase liquid pool morphology on the performance of partitioning tracer tests: evaluation of the equilibrium assumption. Water Resour. Res. 2006, 42, W04408; doi: 10.1029/ 2005WR004074. (16) Mackay, D. M.; Cherry, J. A. Groundwater contamination: Pumpand-treat remediation. Environ. Sci. Technol. 1989, 23, 630636. (17) Anderson, M. R.; Johnson, R. L.; Pankow, J. F. Dissolution of dense chlorinated solvents into groundwater. 3. Modeling contaminant plumes from fingers and pools of solvent. Environ. Sci. Technol. 1992, 26, 901-908. (18) Voudrias, E. A.; Yeh, M.-F. Dissolution of a toluene pool under constant and variable hydraulic gradients with implications for aquifer remediation. Ground Water 1994, 32, 305-311. (19) Seagren, E. A.; Rittmann, B. E.; Valocchi, A. J. An experimental investigation of NAPL pool dissolution enhancement by flushing. J. Contam. Hydrol. 1999, 37, 111-137. (20) Nambi, I. M.; Powers, S. E. NAPL dissolution in heterogeneous systems: an experimental investigation in a simple heterogeneous system. J. Contam. Hydrol. 2000, 44, 161-184. (21) Fetzner, S. Bacterial dehalogenation. Appl. Microbiol. Biotechnol. 1998, 50, 633-657. (22) Isalou, M.; Sleep, B. E.; Liss, S. N. Biodegradation of high concentrations of tetrachloroethene in a continuous flow column system. Environ. Sci. Technol. 1998, 32, 35793585. (23) Nielsen, B. R.; Keasling, J. D. Reductive dechlorination of chlorinated ethenes by a culture enriched from contaminated groundwater. Biotechnol. Bioeng. 1999, 62, 160-165.
(24) Carr, C. S.; Hughes, J. B. Enrichment of high-rate PCE dechlorination and comparative study of lactate, methanol and hydrogen and electron donors to sustain activity. Environ. Sci. Technol. 1998, 32, 1817-1824. (25) Carr, C. S.; Garg, S.; Hughes, J. B. Effect of dechlorinating bacteria on the longevity and composition of PCE-containing nonaqueous phase liquids under equilibrium dissolution conditions. Environ. Sci. Technol. 2000, 34, 1088-1094. (26) Yang, Y.; McCarty, P. L. Biologically enhanced dissolution of tetrachloroethene DNAPL. Environ. Sci. Technol. 2000, 34, 29792984. (27) Cope, N.; Hughes, J. B. Biologically-enhanced removal of PCE from NAPL source zones. Environ. Sci. Technol. 2001, 35, 20142021. (28) Yang, Y.; McCarty, P. Comparison between donor substrates from biologically enhanced tetrachloroethylene DNAPL dissolution. Environ. Sci. Technol. 2002, 36, 3400-3404. (29) Adamson, D. T.; McDade, J. M.; Hughes, J. B. Inoculation of a DNAPL source zone to initiate reductive dechlorination of PCE. Environ. Sci. Technol. 2003, 37, 2525-2533. (30) Song, D. L.; Conrad, M. E.; Sorenson, K. S.; Alvarez-Cohen, L. Stable carbon isotope fractionation during enhanced in situ bioremediation of trichloroethene. Environ. Sci. Technol. 2002, 36, 2262-2268. (31) Horst, J. Successful Biostimulation Treatment of TCE DNAPL. Pollu. Eng. 2006; http://www.pollutioneng.com/CDA/Articles/ Resource_Guide/712033a18670d010VgnVCM100000f932a8c0. (32) Chu, M.; Kitanidis, P. K.; McCarty, P. L. Effects of biomass accumulation on microbially enhanced dissolution of a PCE pool: a numerical simulation. J. Contam. Hydrol. 2003, 65, 79100. (33) Hunt, J. R.; Sitar, N.; Udell, K. S. Nonaqueous phase liquid transport and cleanup: 1. Analysis of mechanism. Water Resour. Res. 1998, 24 (8), 1247-1258. (34) Seagren, E. A.; Rittmann, B. E.; Valocchi, A. J. Bioenhancement of NAPL pool dissolution: experimental evaluation. J. Contam. Hydrol. 2002, 55, 57-85. (35) Edwards, E. A.; Cox, E. E. Field and laboratory studies of sequential anaerobic-aerobic chlorinated solvent biodegradation. In Proceedings of the Fourth International Symposium on In situ and On-Site Bioreclamation, Vol. 3; Leeson, A., Alleman, B. C., Eds.; Battelle Press: Columbus, OH, 1997; pp 261265. (36) Edwards, E. A.; Grbic-Galic, D. Anaerobic degradation of toluene and o-xylene by a methanogenic consortium. Appl. Environ. Microbiol. 1994, 60, 313-322. (37) Ferre, T.; Binley, T.; Geller, J.; Hill, E. H., III; Illangasekare, T. Hydrogeophysical Case Studies at the Lab Scale. In Hydrogeophysics; Y. Rubin, Y., Hubbard, S., Eds.; Springer Verlag: Dordrecht, The Netherlands, 2005; pp 441-466. (38) Mackay, D. M.; Freyberg, D. L.; McCarty, P. L.; Roberts, P. V.; Cherry, J. A. A natural gradient experiment on solute transport in a sand aquifer: I. Approach and overview of plume movement. Water Resour. Res. 1986, 22, 2017-2030. (39) Rehfeldt, K. R.; Boggs, J. M.; Gelhar, L. W. Field study of dispersion in a heterogeneous aquifer. 3: Geostatistical analysis of hydraulic conductivity. Water Resour. Res. 1992, 28, 3309-3324. (40) Corey, A. T. Mechanics of Immiscible Fluids in Porous Media; Water Resources Publications: Highlands Ranch, CO, 1994; 252 pp. (41) Clement, T. P.; Sun, Y.; Hooker, B. S.; Peterson, J. N. Modeling multi-species reactive transport in groundwater aquifers. Ground Water Monit. Rem. 1998, 18, 79-92. (42) Barth, G. R.; Hill, M. H.; Illangasekare, T. H.; Rajaram, H. Predictive modeling of flow and transport in a two-dimensional intermediate-scale heterogeneous porous medium. Water Resour. Res. 2001, 37, 2503-2512. (43) Chu, M.; Kitanidis, P. K.; McCarty, P. L. Possible factors controlling the effectiveness of bioenhanced dissolution of non-aqueous phase tetrachloroethene. Adv. Water Resour. 2004, 27, 601-615.
Received for review April 17, 2006. Revised manuscript received October 16, 2006. Accepted October 17, 2006. ES060922N
VOL. 41, NO. 4, 2007 / ENVIRONMENTAL SCIENCE & TECHNOLOGY
9
1389
