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In situ TCE bioremediation study using electrokinetic cometabolite injection
Waste Management 20 (2000) 279±286
www.elsevier.nl/locate/wasman
In situ TCE bioremediation study using electrokinetic cometabolite injection M.F. Rabbia, B. Clarkc, R.J. Galeb,c,*, E. Ozsu-Acarc, J. Parduea, A. Jacksona a
Department of Civil and Environmental Engineering, Louisiana State University, Baton Rouge, LA 70809-1804, USA b Department of Chemistry, Louisiana State University, Baton Rouge, LA 70809-1804, USA c Electrokinetics Inc., Baton Rouge, LA 70809-1804, USA Accepted 25 October 1999
Abstract The feasibility was evaluated of using electrokinetic injection of benzoic acid cometabolite to enhance the biodegradation of a representative recalcitrant contaminant, trichloroethene (TCE). Whereas in ¯ask studies, sulfate ion alone enhanced TCE (at 6 ppm) degradation rates over those found in the absence of suitable additives, benzoic acid showed enhanced degradation rates for TCE at 6±50 ppm levels. Following injection of benzoic acid cometabolite into a 1 m column of TCE contaminated Loess clay, the TCE ®rst order degradation rate at the periphery was determined to be (0.0390.007) dayÿ1, a value in good agreement with an anaerobic slurry ¯ask tests at 30 C, (0.0470.009) dayÿ1. However, unless the rate of injection of an additive is made compatible with its rate of consumption, these column results and a theoretical model reveal that homogeneous penetration of additive is not achieved. It is cautioned that knowledge of the rate of degradation of a carbon source enhancer (or additive) is critical for engineering its homogeneous injection, whether by hydraulic or electrokinetic methods. These results demonstrate that electrokinetic degradation of recalcitrant wastes may be practical, in particular for those sites whose soil media have low coecients of hydraulic permeability (clay deposits, silty clays, etc.) where traditional pump and treat technology is ineective. # 2000 Elsevier Science Ltd. All rights reserved.
1. Introduction A wide variety of physicochemical treatment technologies are currently available to treat soils contaminated with hazardous materials, including excavation and burial in a chemically secure land®ll, vapor extraction, stabilization and solidi®cation, soil washing, soil ¯ushing, supercritical ¯uid extraction, chemical precipitation, vitri®cation, and incineration [1]. Some of the remediation technologies are not economical and costs can range up to $1500 per cubic yard of soil [2]. Also, many of the available technologies such as pump and treat, do not destroy the hazardous compounds, rather the chemicals are transferred from one phase to another. Since these methods do not completely destroy the contaminants and they are often costly, alternative remedial techniques * Corresponding author. Tel.: +1-225-388-3010; fax: +1-225-3883458. E-mail address: [email protected] (R.J. Gale).
have been and are being investigated. One of the recent technological developments in the ®eld of soil remediation is electrokinetic soil processing. This emerging technology has the potential to remove both inorganic and organic contaminants from soils by using very low DC currents on the order of mA/cm2 of electrode area [3]. Electrokinetic remediation is a technique where electrodes are placed in an open or closed ¯ow arrangement across the soil mass to employ a very low DC electrical current density, or low electrical potential dierence, to transport the species under coupled and/or uncoupled conduction phenomena. These processes result in physicochemical and hydrological changes in the soil matrix [4±7]. Generally, externally supplied ¯uid or groundwater, acts as the conductive medium [5]. The transport of the species under electrical ®elds is in¯uenced by the prevailing electrolysis reactions at the electrodes. This technology has been studied mainly for the remediation of inorganic species; however, it is possible to employ electrokinetics in bioremediation to engender an eective
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level of injected nutrients, electron acceptors/donors, microbes or other process additives in a soil matrix. The process additives can be injected into the system at the electrodes by the electrolysis reactions or by cycling processing ¯uids in the electrode wells. The term bioremediation refers to the complete microbial destruction of contaminants (mineralization process), or transformation of an organic chemical into another form (biotransformation) [8,9]. Electrokinetically enhanced bioremediation (bioelectrokinetic remediation) is being considered as a viable delivery system to enhance bioremediation. Stimulation of microbial populations within a subsurface requires an appropriate carbon source, electron donors/acceptors for energy production, and inorganic nutrients such as nitrogen, phosphorous and some trace metals. Also, proper conditions within the soil environment such as appropriate pH, temperature, moisture content and redox potential are required, which can adequately bolster the growth of desired organisms for a prolonged period. However, there also exist several major technical impediments to the widespread use of biological methods for the treatment of contaminated soils. The full potential of in situ bioremediation is limited due to ignorance of coupled geological, hydrological, physical, chemical, and microbiological processes in the subsurface environment. Many times, the heterogeneous nature of the soil matrix and the contamination itself cannot even be adequately characterized [10]. To develop an optimum and implementable bioremediation design, the site has to be fully characterized. Beside the extent, the type and the degree of contamination, a knowledge of hydrogeological properties are important to design an in situ bioremediation operation [9]. The success of a biological treatment method depends on how eciently and uniformly an engineered system can transport the required process additives, such as electron acceptors/ donors, nutrients, surfactants and cometabolites, into the biologically active zones (BAZs). Consequently, the uniform and ecient introduction of process additives in BAZs has been a bottleneck for successful implementation of in situ bioremediation, e.g. [10,11]. Excessive dosing coupled with the shortcomings of the hydraulically-driven transport processes can result in nutrient rich areas with excessive biological growth (biofouling). Biofouling adversely impacts system implementation due to reduced conductivity by microbial growth plugging the ¯ow paths. This study attempts to assess the electrokinetic enhancement of bioremediation for the anaerobic degradation of trichloroethene (TCE). Its aim is to evaluate and develop in situ bioremediation by electrokinetic injection. The biodegradation process for halogenated aliphatic compounds is quite dierent from that for other common organic contaminants, such as petroleum hydrocarbons. Whereas petroleum hydrocarbons are commonly used as
electron donors in microbial metabolism, there are no known micro-organisms that use halogenated aliphatic compounds as their sole electron (energy) or carbon source. This is because halogenated aliphatic compounds are relatively oxidized and have few available electrons. However, halogenated aliphatic compounds like PCE can be degraded through processes in which they serve as electron acceptors, as well as through various co-metabolic processes [12,13]. Reductive dechlorination, an anaerobic process in which the chlorinated ethene serves as an electron acceptor, is the principal mechanism of bacterial degradation of PCE. Reductive dechlorination is an oxidation-reduction reaction in which electrons are transferred from a donor to a chlorinated hydrocarbon acceptor. The organic substrate is oxidized and the chlorinated hydrocarbon compound is reduced in this process [14]. Reductive dechlorination results in a chlorine being replaced with a hydrogen on the chlorinated hydrocarbon molecule. The process can be carried out on a molecule sequentially, eventually removing all chlorines. However, the pathway and end-products depend upon environmental conditions, including redox status. Under anaerobic conditions, dechlorination generally proceeds in the following sequence [13]: PCE to TCE to DCE to VC to ethene where TCE is trichloroethene, DCE is dichloroethene and VC is vinyl chloride. Three isomers of DCE (1,1-DCE, cis-1,2-DCE and trans-1,2-DCE) may all theoretically be produced in this series of reactions. However, it is reported that cis-1,2-DCE is the most common and 1,1DCE the least common form of DCE formed through reductive dechlorination of TCE [14]. The degradation of vinyl chloride, which may occur through either a reduction or an oxidation reaction, is generally the rate limiting step in the sequence. Vinyl chloride is, therefore, often found to accumulate at contaminated sites. This is of concern because while PCE and TCE are suspected carcinogens, vinyl chloride is a known carcinogen [12]. 2. Experimental 2.1. Shake ¯ask studies of TCE degradation 2.1.1. Kinetic studies Shake ¯ask studies were conducted in order to examine the ability of benzoic acid to stimulate TCE degradation for a representative soil [15]. These studies were conducted anaerobically in 120 ml serum bottles with Te¯on lined stoppers. The Loess soil was a surrogate obtained from Vicksburg, Mississippi. A slurry was produced with the soil by the addition of sterile water
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(autoclaved) and homogenized by vigorous shaking. Bottles received 80 g of slurry (40% water by mass). Three treatments were examined; a killed control (0.2% HgCl2), a live control, and an amended (50 ppm benzoic acid/ wet soil basis). Each treatment was incubated at temperatures of 20, 25, 30 and 45 C. All bottles received TCE to achieve a ®nal concentration of 50 ppm, with the exception of those conducted at 25 C in which the TCE concentration was 6 ppm to investigate if dierent kinetic behaviour occurred at lower levels. It is important to determine if there is a relationship between the degradation rates of TCE and the concentration levels of TCE. In general, some chemical contaminants show zero order degradation rates, while others exhibit ®rst order degradation rates. Knowing the range over which the degradation rate is valid is important in systems which may exhibit highly heterogeneous concentrations. In addition to TCE, all treatments received a mineral salt solution, whose composition is listed in Table 1. An innoculum, which was a 10 g soil sample taken from a Superfund site, Petro Processors, Inc., of known, active TCE degradation capability [16], was added to the soil slurry. In a ®eld site contaminated with TCE, indigenous microbes capable of TCE attack will be present.The innoculum was added to reduce the induction time before degradation commences. This would not be necessary for on-site treatment since the bacterial population would have had time to adapt to the presence of TCE or other contaminants. In addition, microcosm studies were used to investigate the eects of two physical parameters (concentration and temperature), which might be important to this process. TCE was analyzed using a HP 5890 gas chromatograph equipped with an ECD detector and calibrated with commercially prepared standard (Supelco) and the internal standard ``8260 Internal Standards Mix''. 2.2. Column injection studies 2.2.1. Injection studies An electrokinetic injection and a control experiment without electrolysis were compared. The columns were 1 m Table 1 Composition of mineral salt media Compound
mM10ÿ3
KH2PO4 K2HPO4 NaHCO3 MnCl2.4H2O HBO2 ZnCl2 CoCl2.6H2O NiCl2 CuCl2.2H2O Na2MoO4.2H2O
1980 2010 1428 2.5 8 0.4 0.2 0.2 0.2 0.4
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long10.8 cm i.d., packed with Loess soil which was contaminated (nominally) to 100 ppm (wet soil basis). One microliter of microbial innoculum was also added to each kilogram of soil when mixing and packing the columns. Each column had 5 sampling ports along its length. Te¯on endcaps contained electrolyte wells, which were fed by circulatory pumps to a single, isolated source of electrolyte [0.01 M sodium phosphate (monobasic) for the control and 0.01 M sodium phosphate plus 80 ppm benzoic acid for the electrokinetic injection, each with 100 ppm TCE to maintain its concentration during the 25 day injection period]. Mixing of the anolyte and catholyte helps to maintain the pH of the common electrolyte at neutrality and, as necessary, it was adjusted to pH 7 with small quantities of acid/ base. Titanium mesh electrodes were used as anodes and cathodes. The injection period for benzoic acid was 25 days at a constant voltage to provide 1 V/cm gradient across the soil. This period was sucient to allow at least one pore ¯uid of electrolyte to pass from the anode to the cathode. After 25 days, the electrical currents were switched o and the endcaps sealed. Soil samples were taken to establish a baseline for TCE, and at 30 days intervals during the bioremediation stage. The meter columns were used since the fringe eects (in¯uence of reservoirs) are reduced and the distance between sampling ports is reasonable. However, the TCE concentrations were well below the initial (nominal value) of 100 ppm at the start of bioremediation. This loss is due to both volatilization and removal of excess water upon loading the column. Soil core samples were expelled into a tared, Te¯on capped 8 ml vial containing 5.0 ml TCLP grade methanol and stored at 4 C prior to analysis by GC or GC/MS. 3. Results 3.1. Eect of concentration on degradation rates The results of the experiments to determine the role of initial concentration are shown in Fig. 1. This ®gure presents the results of two microcosm experiments using the same soil and techniques with concentrations of TCE varied approximately by an order of magnitude, 6 and 50 ppm, at 25 and 30 C, respectively (the aqueous solubility of TCE at 25 C is approximately 0.11%). Both sets of experiments examined killed control, control, and benzoic amended tests. It is evident that the degradation curves for the benzoic amended treatments are very similar. This implies that under enhanced conditions the degradation of TCE is independent of these initial TCE concentrations. However, it is also apparent that sulfate is as eective as benzoic acid for enhancing the degradation of TCE at the 6 ppm level but the eects of sulfate are less pronounced at higher concentrations
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Fig. 1. Shake ¯ask degradation of TCE at 6 and 50 ppm. Symbols: killed -diamond; control-square; benzoic-triangle; sulfate-circle; benzoic+sulfate-cross.
of TCE. An explanation for these dierences in kinetic behaviour will require a more complete mechanistic study. The killed control shows little loss of TCE supporting the biological role in the transformation of TCE and that volatile loss of TCE is not responsible for the concentration decreases. The dependence or independence of TCE degradation to initial concentration is further illustrated by comparing the rates of TCE degradation for each experiment and treatment. Biodegradation data were ®tted to the following ®rst order degradation equation:
set of microcosm experiments was initiated in order to determine the eects if the temperature were to elevate during the electrokinetic process. These experiments were conducted at 20, 30, and 45 C with/without benzoate amendment. The results are presented in Figs. 2± 4. A degradation rate of 0.04 dayÿ1 indicates a half-life of 17.3 days. Little, if any, degradation was indicated in control treatments again, regardless of temperature. Benzoate amendments clearly stimulated degradation at all temperatures examined. This is further demonstrated in Table 3 which provides the rate constants with their respective standard deviations and r2. As indicated, a temperature of 30 appears to be optimal. However, accelerated degradation is observed over the complete temperature range, 20±45 C, indicating that increased temperature caused by the electrokinetic process may actually stimulate degradation. Temperatures beyond 45 C may inhibit degradation and these temperatures might be reached in the subsurface environment if excessive current densities are used in electrokinetic processing. In this work, the injection period using electrokinetics precedes the bioremediation period and any loss of microbes may result in an extended induction period only.
Table 2 First order degradation rate constants for TCE at 6 ppm (25 ) and 50 ppm (30 ), with/without benzoic acid additive
First order rate (dayÿ1) Standard deviation, s r2
Control 6 ppm
Control 50 ppm
Amended 6 ppm
Amended 50 ppm
0.024 0.005 0.951
0.011 0.009 0.83
0.040 0.005 0.998
0.033 0.005 0.994
C=C0 A B eÿkt where, C=concentration of TCE at time t (days), C0=the initial concentration of TCE, A represents substrate unavailable for degradation, B=total TCE degraded, k=®rst order degradation rate (dayÿ1), s=sample standard deviation, and r2=coecient of determination. A software package (Table Curve) was used to ®t data to the degradation equation and to determine the coecients. The results in Table 2 clearly demonstrates that the degradation rates were very similar for the benzoic acid treatments regardless of TCE concentration. The degradation rates of TCE for the controls were considerably lower than those of the corresponding benzoic acid amended reactions. 3.2. Eect of temperature on degradation rates Some studies [17,19] have indicated that the electrokinetic process causes an increase in soil temperature. A
Fig. 2. Shake ¯ask degradion of TCE at 20 C. Symbols: killed-circle (open); control circle (full); benzoic-square (open); sulfate square (cross); benzoic+sulfate-square (full).
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3.3. Column study of a TCE contaminated loess soil Results are shown in Figs. 5 and 6 for a control and prototype cell experiment. Plots illustrate concentrations of TCE normalized to the concentration present at the end of the electrokinetic injection (25 days), just
Fig. 3. Shake ¯ask degradation of TCE at 30 C. Symbols: killed-circle (open); control circle (full); benzoic-square (open); sulfate square (cross); benzoic+sulfate-square (full).
Fig. 4. Shake ¯ask degradation of TCE at 45 C. Symbols: killed-circle (open); control circle (full); benzoic-square (open); sulfate square (cross); benzoic+sulfate-square (full).
283
prior to the bioremediation period of 90 days. This was done since the experimental column experiences an osmotic ¯ux of anolyte containing benzoic acid. Benzoate anion would also migrate counter to the electroosmotic ¯ow direction. This means that during the initial 25 days, when the electrokinetic current was empowered, TCE concentrations experienced volatilization losses. Once the current was turned o, dierences in the loss of TCE should be due to dierences in the biological reductive dechlorination of TCE, not abiotic processes, since both sealed cells should experience these equally. The control column experienced little loss of TCE from Ports B, C or D, the center of the column, Fig. 5. Ports A and E lost approximately 80% over 120 days. This loss is likely due to volatization, given the fact that it occurred only at the ends of the columns nearest the reservoirs. The benzoic amended column experienced loss of TCE at Ports B, D, and E. Ports A and C exhibited limited losses in TCE concentrations. Ports B and D did experience loss of TCE, which would be less likely from abiotic loss and more likely due to biological activity. As mentioned, port B in the control core experienced no loss. Only Port C in the benzoic amended cell exhibited little loss of TCE. The other two Ports B and D did exhibit signi®cant loss of TCE, during the injection period, which is unlikely to be contributable to
Fig. 5. TCE analyses at 30 day intervals in control column, normalized to start of bioremediation period.
Table 3 First order degradation rate constants of TCE at 20, 30, and 45 C, with/without benzoic acid, 50 ppm
First order rate (dayÿ1) Standard deviation, s r2
T 20 C w/o
T 20 C w
T 30 C w/o
T 30 C w
T 45 C w/o
T 45 C w
0.01 0.009 0.83
0.03 0.005 0.994
0.014 0.022 0.82
0.047 0.009 0.98
0.007 0.006 0.982
0.032 0.006 0.99
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abiotic processes since no such loss was observed in the control cell. However, in order to be conservative in our analysis, it was decided that only Ports in which biological metabolic daughter products of TCE (cis or trans DCE, and vinyl chloride) appeared would be labeled as experiencing biological degradation of TCE. The strongest evidence for degradation is the appearance of lower chlorinated metabolites of TCE dechlorination, primarily cis-dichloroethylene (cis-DCE) and transdichloroethylene (trans-DCE), at Port D in the columns with electrokinetic injection of benzoate ion. These two breakdown products have been shown to be the major biological degradation products of TCE degradation. These breakdown products were observed equivalent to 25% of the available TCE mass. TCE degradation
appears to be minimal in the unamended cell and only trace levels of its metabolites were detected. Fig. 7 shows a comparison of the control and benzoate amended treatments for the one-meter column study and the microcosm shake ¯ask study. The apparent loss rates of (0.0390.007) dayÿ1 (r2=0.998) for columns with benzoate delivered by electrokinetic injection is much greater than the column that received only the electrokinetic current, which had apparent loss rates of (0.0020.001) dayÿ1 (r2=0.51). This rate of TCE degradation is virtually identical to the batch microcosm studies (0.0380.005 dayÿ1), (Table 2). The determined TCE degradation rate constant for the benzoate amended columns is greater than the rate constants reported in the literature, which range from (0.0027 to 0.011 dayÿ1) but no rate constants for benzoate-enhanced degradation were available for direct comparison. 3.4. Theoretical analysis of additive injection A computer simulation was used to model the injection homogeneity of a carbon source (cometabolite) for microbial degradation enhancement. It is assumed the additive (e.g., benzoate ion) was injected electroosmotically at 4 cm/day from a 100 ppm source well, with a homogeneous ®rst order rate constant for microbial degradation k 0:1 dayÿ1 (one week half life approx). With a simple BASIC program and using ®nite elements T 6 h, L 1:00 cm, the steady-state concentration pro®les of additive to a penetration depth of 1 m can be computed from the expression,
Fig. 6. TCE analyses at 30 day intervals in electrokinetic column, normalized to start of bioremediation period.
A N A N ÿ 1 exp ÿk:T
Fig. 7. Comparison of TCE losses in control and benzoic acid amended cells. Symbols: controls-square (open); benzoic-circle (open).
Fig. 8. Simulation of additive penetration by ®rst order homogeneous microbial degradation at k 0:1, 1 and 10 dayÿ1 rate constants.
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The ®rst, second and third elements are 97.5, 95.1, 92.8,.......ppm and it is interesting to note, that even with the slowest constant selected, the concentration at 1 m depth is an order of magnitude smaller than the source, Fig. 8. This demonstrates a fundamental limitation to injection, either by hydraulic or electrokinetic means, when there is sucient microbial populations to cause ®rst order reaction. If an additive is readily consummed then the microbes at the peripheries will prevent its ingress to the bulk and possibly continued injection will result in biofouling. Injection systems must be engineered with a knowledge of the kinetics of the carbon source degradation. It is apparent that slowly degrading or slowly released initiators must be found if bioremediation is to be enhanced through additive dosing. 4. Conclusions This study has explored the feasibility of using electrokinetic injection of nutrient ions and amendments to enhance the biodegradation of TCE. Benzoic acid, a bioremediation cometabolite for TCE degradation was electrokinetically injected by both electroosmosis (neutral benzoic acid) and ion injection (the benzoate anion). However, the rate of degradation of this additive requires additional study since microbial activity may be preventing its homogeneous injection. Following injection of this cometabolite in a 1 m column, the TCE ®rst order degradation rate close to the periphery was determined to be (0.0390.007) dayÿ1, a value in good agreement with that determined in an anaerobic slurry ¯ask test at 30 C (0.0470.009) day ÿ1. These experiments demonstrate that electrokinetic injection to engineer degradation of recalcitrant hydrocarbons, or other dicult to degrade contaminants is feasible in principle. However, it is cautioned that knowledge of the rates of degradation of carbon source enhancers are critical for ensuring their homogeneous distribution. Fundamental studies are needed of the consumption rates of nutrients and carbon sources by microbes in order to engineer and optimize injection protocols. To achieve homogenous injection, the penetration rate must exceed the local degradation rate of the carbon source. It is hoped that a ®eld pilot-scale or small volume TCE spill be remediated by bioelectrokinetic remediation to hasten the practical utility of this technology for full-scale site applications. Loo and Chilinger [18] have discussed the results of some site studies using bioelectrokinetic remediation to degrade contaminants. Apparently, these demonstrations were successful but details of the precise treatment schemes are not available. A cost quoted for remediation of 100,000 cubic yards of soil contaminated with chlorinated solvents at Northrop ESP, Anaheim, CA is stated to be less than $20/cubic yard.
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Acknowledgements Electrokinetics Inc. is appreciative of the USEPA Oce of Research and Development, Cincinnati, particularly the Project Ocer Randy Parker, for support for this work. Studies were in part funded by the LEQSF Program, Board of Regents, LSU, through the Industrial Ties Sub-program. References [1] U.S. EPA. Technology screen guide for treatment of CERCLA soils and sludges. Publ. EPA/540/2-88/004, Washington DC, 1988. [2] Acar YB. Electrokinetic soil processing; a review of the state of the art, ASCE Grouting Conference, February 1992, ASCE Special Publication No. 30, vol. 2, p. 1420±1432. [3] Acar YB, Gale RJ, Alshawabkeh AN, Marks R, Puppala S, Bricka M et al. Electrokinetic remediation: basics and technology status. Journal of Haz Mat 1995;40(2):117±37. [4] Acar YB, Alshawabkeh AN, Gale RJ. Fundamental aspects of extracting species from soils by electrokinetics. Waste Management 1993;12(3):1410±21. [5] Acar YB, Alshawabkeh AN. Principles of electrokinetic remediation. Environmental Science and Technology 1993;27(13):2638± 47. [6] Runnells DD, Wahli C. In-situ electromigration as a method for removing sulfate, metals and other contaminants from ground water. Ground Water Monitoring Review 1993;11(3):121. [7] Pamukcu S, Whittle JK. Electrokinetic removal of selected heavy metals from soil. Environ Progress 1993;11(3):241±50. [8] Slater HJ, Lovatt D. Biodegradation and signi®cance of microbial communities. In: Gibson DT, editor. Microbial degradation of organic compounds. New York: Marcel Dekker, 1984. p. 439±85. [9] Skladany GJ, Metting FB. Bioremedation of contaminated soils. In: Metting Jr. FB, editor. Soil microbial ecology. New York: Marcel Dekker Inc, 1993. p. 483±513. [10] Su¯ita JM, Sewell GW. Anaerobic biotransformation of contaminants in the subsurface. EPA, Robert S. Kerr Envir. Laboratory, Ada, OK EPA/600/M-90/024, 1991. [11] Zappi M, Gunnison D, Pennington J, Cullinane J, Teeter CL, Brannon JM, Myers T. Technical approach for in situ biological treatment research: bench-scale experiments, Tech. Rpt. No. IRRP-93-3, Aug. 1993, US Army Corps of Engineers, Waterways Experiment Station, 1993. [12] Chapelle FH. Ground-water microbiology and geochemistry. New York: John Wiley & Sons, 1993. [13] Chapelle FH. Identifying redox conditions that favor the natural attenuation of chlorinated ethenes in contaminated ground-water systems. In Symposium on Natural Attenuation of Chlorinated Organics in Ground Water, USEPA/540/R-96/509, 1996. [14] Wiedemeier, TH, Swanson MA, Moutoux DE. Overview of the technical protocol for natural attenuation of chlorinated aliphatic hydrocarbons in groundwater under development for the U.S. air force center for environmental excellence. In Symposium on Natural Attenuation of Chlorinated Organics in Groundwater, USEPA/540/R-96/509, 1996. [15] Beeman RE, Howell JE, Shoemaker SH, Salazar EA, Buttram JR. A ®eld evaluation of in situ microbial reductive dehalogenation by the biotransformation of chlorinated ethenes. In:Hinchee R, Leeson A, Semprini L, Ong SK, editor. Bioremediation of chlorinated and PAH compounds. Lewis, 1994. p. 14±27. [16] Trail, K. Master's thesis, May 1998, Biostimulation versus intrinsic bioremediation of TCE in a contaminated aquifer at a
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superfund site. Dept. Of Civil and Environ. Engr., Louisiana State University, 1998. [17] Ho SV, Sheridan PW, Athmer CJ. Electrokinetic aspects of the Lasagna process for in situ soil remediation. Extended abstract. Emerging Technologies in hazardous waste management VI, I. & E.C. Div. ACS, 19±21 September, D.W. Tedder (Ed), 1994.
[18] Loo WW, Chilingar GV. Advances in the electrokinetic treatment of hazardous wastes in soil and groundwater. Remediation Management, Second Quarter 1997, p. 39±41. [19] Athmer CJ, Ho SV, Hughes BM. Results of the Lasagna phase II, Field Experiment. Extended abstract. Emerging technologies in hazardous waste management IX, 15±17, September 1997.
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In situ TCE bioremediation study using electrokinetic cometabolite injection M.F. Rabbia, B. Clarkc, R.J. Galeb,c,*, E. Ozsu-Acarc, J. Parduea, A. Jacksona a
Department of Civil and Environmental Engineering, Louisiana State University, Baton Rouge, LA 70809-1804, USA b Department of Chemistry, Louisiana State University, Baton Rouge, LA 70809-1804, USA c Electrokinetics Inc., Baton Rouge, LA 70809-1804, USA Accepted 25 October 1999
Abstract The feasibility was evaluated of using electrokinetic injection of benzoic acid cometabolite to enhance the biodegradation of a representative recalcitrant contaminant, trichloroethene (TCE). Whereas in ¯ask studies, sulfate ion alone enhanced TCE (at 6 ppm) degradation rates over those found in the absence of suitable additives, benzoic acid showed enhanced degradation rates for TCE at 6±50 ppm levels. Following injection of benzoic acid cometabolite into a 1 m column of TCE contaminated Loess clay, the TCE ®rst order degradation rate at the periphery was determined to be (0.0390.007) dayÿ1, a value in good agreement with an anaerobic slurry ¯ask tests at 30 C, (0.0470.009) dayÿ1. However, unless the rate of injection of an additive is made compatible with its rate of consumption, these column results and a theoretical model reveal that homogeneous penetration of additive is not achieved. It is cautioned that knowledge of the rate of degradation of a carbon source enhancer (or additive) is critical for engineering its homogeneous injection, whether by hydraulic or electrokinetic methods. These results demonstrate that electrokinetic degradation of recalcitrant wastes may be practical, in particular for those sites whose soil media have low coecients of hydraulic permeability (clay deposits, silty clays, etc.) where traditional pump and treat technology is ineective. # 2000 Elsevier Science Ltd. All rights reserved.
1. Introduction A wide variety of physicochemical treatment technologies are currently available to treat soils contaminated with hazardous materials, including excavation and burial in a chemically secure land®ll, vapor extraction, stabilization and solidi®cation, soil washing, soil ¯ushing, supercritical ¯uid extraction, chemical precipitation, vitri®cation, and incineration [1]. Some of the remediation technologies are not economical and costs can range up to $1500 per cubic yard of soil [2]. Also, many of the available technologies such as pump and treat, do not destroy the hazardous compounds, rather the chemicals are transferred from one phase to another. Since these methods do not completely destroy the contaminants and they are often costly, alternative remedial techniques * Corresponding author. Tel.: +1-225-388-3010; fax: +1-225-3883458. E-mail address: [email protected] (R.J. Gale).
have been and are being investigated. One of the recent technological developments in the ®eld of soil remediation is electrokinetic soil processing. This emerging technology has the potential to remove both inorganic and organic contaminants from soils by using very low DC currents on the order of mA/cm2 of electrode area [3]. Electrokinetic remediation is a technique where electrodes are placed in an open or closed ¯ow arrangement across the soil mass to employ a very low DC electrical current density, or low electrical potential dierence, to transport the species under coupled and/or uncoupled conduction phenomena. These processes result in physicochemical and hydrological changes in the soil matrix [4±7]. Generally, externally supplied ¯uid or groundwater, acts as the conductive medium [5]. The transport of the species under electrical ®elds is in¯uenced by the prevailing electrolysis reactions at the electrodes. This technology has been studied mainly for the remediation of inorganic species; however, it is possible to employ electrokinetics in bioremediation to engender an eective
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level of injected nutrients, electron acceptors/donors, microbes or other process additives in a soil matrix. The process additives can be injected into the system at the electrodes by the electrolysis reactions or by cycling processing ¯uids in the electrode wells. The term bioremediation refers to the complete microbial destruction of contaminants (mineralization process), or transformation of an organic chemical into another form (biotransformation) [8,9]. Electrokinetically enhanced bioremediation (bioelectrokinetic remediation) is being considered as a viable delivery system to enhance bioremediation. Stimulation of microbial populations within a subsurface requires an appropriate carbon source, electron donors/acceptors for energy production, and inorganic nutrients such as nitrogen, phosphorous and some trace metals. Also, proper conditions within the soil environment such as appropriate pH, temperature, moisture content and redox potential are required, which can adequately bolster the growth of desired organisms for a prolonged period. However, there also exist several major technical impediments to the widespread use of biological methods for the treatment of contaminated soils. The full potential of in situ bioremediation is limited due to ignorance of coupled geological, hydrological, physical, chemical, and microbiological processes in the subsurface environment. Many times, the heterogeneous nature of the soil matrix and the contamination itself cannot even be adequately characterized [10]. To develop an optimum and implementable bioremediation design, the site has to be fully characterized. Beside the extent, the type and the degree of contamination, a knowledge of hydrogeological properties are important to design an in situ bioremediation operation [9]. The success of a biological treatment method depends on how eciently and uniformly an engineered system can transport the required process additives, such as electron acceptors/ donors, nutrients, surfactants and cometabolites, into the biologically active zones (BAZs). Consequently, the uniform and ecient introduction of process additives in BAZs has been a bottleneck for successful implementation of in situ bioremediation, e.g. [10,11]. Excessive dosing coupled with the shortcomings of the hydraulically-driven transport processes can result in nutrient rich areas with excessive biological growth (biofouling). Biofouling adversely impacts system implementation due to reduced conductivity by microbial growth plugging the ¯ow paths. This study attempts to assess the electrokinetic enhancement of bioremediation for the anaerobic degradation of trichloroethene (TCE). Its aim is to evaluate and develop in situ bioremediation by electrokinetic injection. The biodegradation process for halogenated aliphatic compounds is quite dierent from that for other common organic contaminants, such as petroleum hydrocarbons. Whereas petroleum hydrocarbons are commonly used as
electron donors in microbial metabolism, there are no known micro-organisms that use halogenated aliphatic compounds as their sole electron (energy) or carbon source. This is because halogenated aliphatic compounds are relatively oxidized and have few available electrons. However, halogenated aliphatic compounds like PCE can be degraded through processes in which they serve as electron acceptors, as well as through various co-metabolic processes [12,13]. Reductive dechlorination, an anaerobic process in which the chlorinated ethene serves as an electron acceptor, is the principal mechanism of bacterial degradation of PCE. Reductive dechlorination is an oxidation-reduction reaction in which electrons are transferred from a donor to a chlorinated hydrocarbon acceptor. The organic substrate is oxidized and the chlorinated hydrocarbon compound is reduced in this process [14]. Reductive dechlorination results in a chlorine being replaced with a hydrogen on the chlorinated hydrocarbon molecule. The process can be carried out on a molecule sequentially, eventually removing all chlorines. However, the pathway and end-products depend upon environmental conditions, including redox status. Under anaerobic conditions, dechlorination generally proceeds in the following sequence [13]: PCE to TCE to DCE to VC to ethene where TCE is trichloroethene, DCE is dichloroethene and VC is vinyl chloride. Three isomers of DCE (1,1-DCE, cis-1,2-DCE and trans-1,2-DCE) may all theoretically be produced in this series of reactions. However, it is reported that cis-1,2-DCE is the most common and 1,1DCE the least common form of DCE formed through reductive dechlorination of TCE [14]. The degradation of vinyl chloride, which may occur through either a reduction or an oxidation reaction, is generally the rate limiting step in the sequence. Vinyl chloride is, therefore, often found to accumulate at contaminated sites. This is of concern because while PCE and TCE are suspected carcinogens, vinyl chloride is a known carcinogen [12]. 2. Experimental 2.1. Shake ¯ask studies of TCE degradation 2.1.1. Kinetic studies Shake ¯ask studies were conducted in order to examine the ability of benzoic acid to stimulate TCE degradation for a representative soil [15]. These studies were conducted anaerobically in 120 ml serum bottles with Te¯on lined stoppers. The Loess soil was a surrogate obtained from Vicksburg, Mississippi. A slurry was produced with the soil by the addition of sterile water
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(autoclaved) and homogenized by vigorous shaking. Bottles received 80 g of slurry (40% water by mass). Three treatments were examined; a killed control (0.2% HgCl2), a live control, and an amended (50 ppm benzoic acid/ wet soil basis). Each treatment was incubated at temperatures of 20, 25, 30 and 45 C. All bottles received TCE to achieve a ®nal concentration of 50 ppm, with the exception of those conducted at 25 C in which the TCE concentration was 6 ppm to investigate if dierent kinetic behaviour occurred at lower levels. It is important to determine if there is a relationship between the degradation rates of TCE and the concentration levels of TCE. In general, some chemical contaminants show zero order degradation rates, while others exhibit ®rst order degradation rates. Knowing the range over which the degradation rate is valid is important in systems which may exhibit highly heterogeneous concentrations. In addition to TCE, all treatments received a mineral salt solution, whose composition is listed in Table 1. An innoculum, which was a 10 g soil sample taken from a Superfund site, Petro Processors, Inc., of known, active TCE degradation capability [16], was added to the soil slurry. In a ®eld site contaminated with TCE, indigenous microbes capable of TCE attack will be present.The innoculum was added to reduce the induction time before degradation commences. This would not be necessary for on-site treatment since the bacterial population would have had time to adapt to the presence of TCE or other contaminants. In addition, microcosm studies were used to investigate the eects of two physical parameters (concentration and temperature), which might be important to this process. TCE was analyzed using a HP 5890 gas chromatograph equipped with an ECD detector and calibrated with commercially prepared standard (Supelco) and the internal standard ``8260 Internal Standards Mix''. 2.2. Column injection studies 2.2.1. Injection studies An electrokinetic injection and a control experiment without electrolysis were compared. The columns were 1 m Table 1 Composition of mineral salt media Compound
mM10ÿ3
KH2PO4 K2HPO4 NaHCO3 MnCl2.4H2O HBO2 ZnCl2 CoCl2.6H2O NiCl2 CuCl2.2H2O Na2MoO4.2H2O
1980 2010 1428 2.5 8 0.4 0.2 0.2 0.2 0.4
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long10.8 cm i.d., packed with Loess soil which was contaminated (nominally) to 100 ppm (wet soil basis). One microliter of microbial innoculum was also added to each kilogram of soil when mixing and packing the columns. Each column had 5 sampling ports along its length. Te¯on endcaps contained electrolyte wells, which were fed by circulatory pumps to a single, isolated source of electrolyte [0.01 M sodium phosphate (monobasic) for the control and 0.01 M sodium phosphate plus 80 ppm benzoic acid for the electrokinetic injection, each with 100 ppm TCE to maintain its concentration during the 25 day injection period]. Mixing of the anolyte and catholyte helps to maintain the pH of the common electrolyte at neutrality and, as necessary, it was adjusted to pH 7 with small quantities of acid/ base. Titanium mesh electrodes were used as anodes and cathodes. The injection period for benzoic acid was 25 days at a constant voltage to provide 1 V/cm gradient across the soil. This period was sucient to allow at least one pore ¯uid of electrolyte to pass from the anode to the cathode. After 25 days, the electrical currents were switched o and the endcaps sealed. Soil samples were taken to establish a baseline for TCE, and at 30 days intervals during the bioremediation stage. The meter columns were used since the fringe eects (in¯uence of reservoirs) are reduced and the distance between sampling ports is reasonable. However, the TCE concentrations were well below the initial (nominal value) of 100 ppm at the start of bioremediation. This loss is due to both volatilization and removal of excess water upon loading the column. Soil core samples were expelled into a tared, Te¯on capped 8 ml vial containing 5.0 ml TCLP grade methanol and stored at 4 C prior to analysis by GC or GC/MS. 3. Results 3.1. Eect of concentration on degradation rates The results of the experiments to determine the role of initial concentration are shown in Fig. 1. This ®gure presents the results of two microcosm experiments using the same soil and techniques with concentrations of TCE varied approximately by an order of magnitude, 6 and 50 ppm, at 25 and 30 C, respectively (the aqueous solubility of TCE at 25 C is approximately 0.11%). Both sets of experiments examined killed control, control, and benzoic amended tests. It is evident that the degradation curves for the benzoic amended treatments are very similar. This implies that under enhanced conditions the degradation of TCE is independent of these initial TCE concentrations. However, it is also apparent that sulfate is as eective as benzoic acid for enhancing the degradation of TCE at the 6 ppm level but the eects of sulfate are less pronounced at higher concentrations
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Fig. 1. Shake ¯ask degradation of TCE at 6 and 50 ppm. Symbols: killed -diamond; control-square; benzoic-triangle; sulfate-circle; benzoic+sulfate-cross.
of TCE. An explanation for these dierences in kinetic behaviour will require a more complete mechanistic study. The killed control shows little loss of TCE supporting the biological role in the transformation of TCE and that volatile loss of TCE is not responsible for the concentration decreases. The dependence or independence of TCE degradation to initial concentration is further illustrated by comparing the rates of TCE degradation for each experiment and treatment. Biodegradation data were ®tted to the following ®rst order degradation equation:
set of microcosm experiments was initiated in order to determine the eects if the temperature were to elevate during the electrokinetic process. These experiments were conducted at 20, 30, and 45 C with/without benzoate amendment. The results are presented in Figs. 2± 4. A degradation rate of 0.04 dayÿ1 indicates a half-life of 17.3 days. Little, if any, degradation was indicated in control treatments again, regardless of temperature. Benzoate amendments clearly stimulated degradation at all temperatures examined. This is further demonstrated in Table 3 which provides the rate constants with their respective standard deviations and r2. As indicated, a temperature of 30 appears to be optimal. However, accelerated degradation is observed over the complete temperature range, 20±45 C, indicating that increased temperature caused by the electrokinetic process may actually stimulate degradation. Temperatures beyond 45 C may inhibit degradation and these temperatures might be reached in the subsurface environment if excessive current densities are used in electrokinetic processing. In this work, the injection period using electrokinetics precedes the bioremediation period and any loss of microbes may result in an extended induction period only.
Table 2 First order degradation rate constants for TCE at 6 ppm (25 ) and 50 ppm (30 ), with/without benzoic acid additive
First order rate (dayÿ1) Standard deviation, s r2
Control 6 ppm
Control 50 ppm
Amended 6 ppm
Amended 50 ppm
0.024 0.005 0.951
0.011 0.009 0.83
0.040 0.005 0.998
0.033 0.005 0.994
C=C0 A B eÿkt where, C=concentration of TCE at time t (days), C0=the initial concentration of TCE, A represents substrate unavailable for degradation, B=total TCE degraded, k=®rst order degradation rate (dayÿ1), s=sample standard deviation, and r2=coecient of determination. A software package (Table Curve) was used to ®t data to the degradation equation and to determine the coecients. The results in Table 2 clearly demonstrates that the degradation rates were very similar for the benzoic acid treatments regardless of TCE concentration. The degradation rates of TCE for the controls were considerably lower than those of the corresponding benzoic acid amended reactions. 3.2. Eect of temperature on degradation rates Some studies [17,19] have indicated that the electrokinetic process causes an increase in soil temperature. A
Fig. 2. Shake ¯ask degradion of TCE at 20 C. Symbols: killed-circle (open); control circle (full); benzoic-square (open); sulfate square (cross); benzoic+sulfate-square (full).
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3.3. Column study of a TCE contaminated loess soil Results are shown in Figs. 5 and 6 for a control and prototype cell experiment. Plots illustrate concentrations of TCE normalized to the concentration present at the end of the electrokinetic injection (25 days), just
Fig. 3. Shake ¯ask degradation of TCE at 30 C. Symbols: killed-circle (open); control circle (full); benzoic-square (open); sulfate square (cross); benzoic+sulfate-square (full).
Fig. 4. Shake ¯ask degradation of TCE at 45 C. Symbols: killed-circle (open); control circle (full); benzoic-square (open); sulfate square (cross); benzoic+sulfate-square (full).
283
prior to the bioremediation period of 90 days. This was done since the experimental column experiences an osmotic ¯ux of anolyte containing benzoic acid. Benzoate anion would also migrate counter to the electroosmotic ¯ow direction. This means that during the initial 25 days, when the electrokinetic current was empowered, TCE concentrations experienced volatilization losses. Once the current was turned o, dierences in the loss of TCE should be due to dierences in the biological reductive dechlorination of TCE, not abiotic processes, since both sealed cells should experience these equally. The control column experienced little loss of TCE from Ports B, C or D, the center of the column, Fig. 5. Ports A and E lost approximately 80% over 120 days. This loss is likely due to volatization, given the fact that it occurred only at the ends of the columns nearest the reservoirs. The benzoic amended column experienced loss of TCE at Ports B, D, and E. Ports A and C exhibited limited losses in TCE concentrations. Ports B and D did experience loss of TCE, which would be less likely from abiotic loss and more likely due to biological activity. As mentioned, port B in the control core experienced no loss. Only Port C in the benzoic amended cell exhibited little loss of TCE. The other two Ports B and D did exhibit signi®cant loss of TCE, during the injection period, which is unlikely to be contributable to
Fig. 5. TCE analyses at 30 day intervals in control column, normalized to start of bioremediation period.
Table 3 First order degradation rate constants of TCE at 20, 30, and 45 C, with/without benzoic acid, 50 ppm
First order rate (dayÿ1) Standard deviation, s r2
T 20 C w/o
T 20 C w
T 30 C w/o
T 30 C w
T 45 C w/o
T 45 C w
0.01 0.009 0.83
0.03 0.005 0.994
0.014 0.022 0.82
0.047 0.009 0.98
0.007 0.006 0.982
0.032 0.006 0.99
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abiotic processes since no such loss was observed in the control cell. However, in order to be conservative in our analysis, it was decided that only Ports in which biological metabolic daughter products of TCE (cis or trans DCE, and vinyl chloride) appeared would be labeled as experiencing biological degradation of TCE. The strongest evidence for degradation is the appearance of lower chlorinated metabolites of TCE dechlorination, primarily cis-dichloroethylene (cis-DCE) and transdichloroethylene (trans-DCE), at Port D in the columns with electrokinetic injection of benzoate ion. These two breakdown products have been shown to be the major biological degradation products of TCE degradation. These breakdown products were observed equivalent to 25% of the available TCE mass. TCE degradation
appears to be minimal in the unamended cell and only trace levels of its metabolites were detected. Fig. 7 shows a comparison of the control and benzoate amended treatments for the one-meter column study and the microcosm shake ¯ask study. The apparent loss rates of (0.0390.007) dayÿ1 (r2=0.998) for columns with benzoate delivered by electrokinetic injection is much greater than the column that received only the electrokinetic current, which had apparent loss rates of (0.0020.001) dayÿ1 (r2=0.51). This rate of TCE degradation is virtually identical to the batch microcosm studies (0.0380.005 dayÿ1), (Table 2). The determined TCE degradation rate constant for the benzoate amended columns is greater than the rate constants reported in the literature, which range from (0.0027 to 0.011 dayÿ1) but no rate constants for benzoate-enhanced degradation were available for direct comparison. 3.4. Theoretical analysis of additive injection A computer simulation was used to model the injection homogeneity of a carbon source (cometabolite) for microbial degradation enhancement. It is assumed the additive (e.g., benzoate ion) was injected electroosmotically at 4 cm/day from a 100 ppm source well, with a homogeneous ®rst order rate constant for microbial degradation k 0:1 dayÿ1 (one week half life approx). With a simple BASIC program and using ®nite elements T 6 h, L 1:00 cm, the steady-state concentration pro®les of additive to a penetration depth of 1 m can be computed from the expression,
Fig. 6. TCE analyses at 30 day intervals in electrokinetic column, normalized to start of bioremediation period.
A N A N ÿ 1 exp ÿk:T
Fig. 7. Comparison of TCE losses in control and benzoic acid amended cells. Symbols: controls-square (open); benzoic-circle (open).
Fig. 8. Simulation of additive penetration by ®rst order homogeneous microbial degradation at k 0:1, 1 and 10 dayÿ1 rate constants.
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The ®rst, second and third elements are 97.5, 95.1, 92.8,.......ppm and it is interesting to note, that even with the slowest constant selected, the concentration at 1 m depth is an order of magnitude smaller than the source, Fig. 8. This demonstrates a fundamental limitation to injection, either by hydraulic or electrokinetic means, when there is sucient microbial populations to cause ®rst order reaction. If an additive is readily consummed then the microbes at the peripheries will prevent its ingress to the bulk and possibly continued injection will result in biofouling. Injection systems must be engineered with a knowledge of the kinetics of the carbon source degradation. It is apparent that slowly degrading or slowly released initiators must be found if bioremediation is to be enhanced through additive dosing. 4. Conclusions This study has explored the feasibility of using electrokinetic injection of nutrient ions and amendments to enhance the biodegradation of TCE. Benzoic acid, a bioremediation cometabolite for TCE degradation was electrokinetically injected by both electroosmosis (neutral benzoic acid) and ion injection (the benzoate anion). However, the rate of degradation of this additive requires additional study since microbial activity may be preventing its homogeneous injection. Following injection of this cometabolite in a 1 m column, the TCE ®rst order degradation rate close to the periphery was determined to be (0.0390.007) dayÿ1, a value in good agreement with that determined in an anaerobic slurry ¯ask test at 30 C (0.0470.009) day ÿ1. These experiments demonstrate that electrokinetic injection to engineer degradation of recalcitrant hydrocarbons, or other dicult to degrade contaminants is feasible in principle. However, it is cautioned that knowledge of the rates of degradation of carbon source enhancers are critical for ensuring their homogeneous distribution. Fundamental studies are needed of the consumption rates of nutrients and carbon sources by microbes in order to engineer and optimize injection protocols. To achieve homogenous injection, the penetration rate must exceed the local degradation rate of the carbon source. It is hoped that a ®eld pilot-scale or small volume TCE spill be remediated by bioelectrokinetic remediation to hasten the practical utility of this technology for full-scale site applications. Loo and Chilinger [18] have discussed the results of some site studies using bioelectrokinetic remediation to degrade contaminants. Apparently, these demonstrations were successful but details of the precise treatment schemes are not available. A cost quoted for remediation of 100,000 cubic yards of soil contaminated with chlorinated solvents at Northrop ESP, Anaheim, CA is stated to be less than $20/cubic yard.
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Acknowledgements Electrokinetics Inc. is appreciative of the USEPA Oce of Research and Development, Cincinnati, particularly the Project Ocer Randy Parker, for support for this work. Studies were in part funded by the LEQSF Program, Board of Regents, LSU, through the Industrial Ties Sub-program. References [1] U.S. EPA. Technology screen guide for treatment of CERCLA soils and sludges. Publ. EPA/540/2-88/004, Washington DC, 1988. [2] Acar YB. Electrokinetic soil processing; a review of the state of the art, ASCE Grouting Conference, February 1992, ASCE Special Publication No. 30, vol. 2, p. 1420±1432. [3] Acar YB, Gale RJ, Alshawabkeh AN, Marks R, Puppala S, Bricka M et al. Electrokinetic remediation: basics and technology status. Journal of Haz Mat 1995;40(2):117±37. [4] Acar YB, Alshawabkeh AN, Gale RJ. Fundamental aspects of extracting species from soils by electrokinetics. Waste Management 1993;12(3):1410±21. [5] Acar YB, Alshawabkeh AN. Principles of electrokinetic remediation. Environmental Science and Technology 1993;27(13):2638± 47. [6] Runnells DD, Wahli C. In-situ electromigration as a method for removing sulfate, metals and other contaminants from ground water. Ground Water Monitoring Review 1993;11(3):121. [7] Pamukcu S, Whittle JK. Electrokinetic removal of selected heavy metals from soil. Environ Progress 1993;11(3):241±50. [8] Slater HJ, Lovatt D. Biodegradation and signi®cance of microbial communities. In: Gibson DT, editor. Microbial degradation of organic compounds. New York: Marcel Dekker, 1984. p. 439±85. [9] Skladany GJ, Metting FB. Bioremedation of contaminated soils. In: Metting Jr. FB, editor. Soil microbial ecology. New York: Marcel Dekker Inc, 1993. p. 483±513. [10] Su¯ita JM, Sewell GW. Anaerobic biotransformation of contaminants in the subsurface. EPA, Robert S. Kerr Envir. Laboratory, Ada, OK EPA/600/M-90/024, 1991. [11] Zappi M, Gunnison D, Pennington J, Cullinane J, Teeter CL, Brannon JM, Myers T. Technical approach for in situ biological treatment research: bench-scale experiments, Tech. Rpt. No. IRRP-93-3, Aug. 1993, US Army Corps of Engineers, Waterways Experiment Station, 1993. [12] Chapelle FH. Ground-water microbiology and geochemistry. New York: John Wiley & Sons, 1993. [13] Chapelle FH. Identifying redox conditions that favor the natural attenuation of chlorinated ethenes in contaminated ground-water systems. In Symposium on Natural Attenuation of Chlorinated Organics in Ground Water, USEPA/540/R-96/509, 1996. [14] Wiedemeier, TH, Swanson MA, Moutoux DE. Overview of the technical protocol for natural attenuation of chlorinated aliphatic hydrocarbons in groundwater under development for the U.S. air force center for environmental excellence. In Symposium on Natural Attenuation of Chlorinated Organics in Groundwater, USEPA/540/R-96/509, 1996. [15] Beeman RE, Howell JE, Shoemaker SH, Salazar EA, Buttram JR. A ®eld evaluation of in situ microbial reductive dehalogenation by the biotransformation of chlorinated ethenes. In:Hinchee R, Leeson A, Semprini L, Ong SK, editor. Bioremediation of chlorinated and PAH compounds. Lewis, 1994. p. 14±27. [16] Trail, K. Master's thesis, May 1998, Biostimulation versus intrinsic bioremediation of TCE in a contaminated aquifer at a
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superfund site. Dept. Of Civil and Environ. Engr., Louisiana State University, 1998. [17] Ho SV, Sheridan PW, Athmer CJ. Electrokinetic aspects of the Lasagna process for in situ soil remediation. Extended abstract. Emerging Technologies in hazardous waste management VI, I. & E.C. Div. ACS, 19±21 September, D.W. Tedder (Ed), 1994.
[18] Loo WW, Chilingar GV. Advances in the electrokinetic treatment of hazardous wastes in soil and groundwater. Remediation Management, Second Quarter 1997, p. 39±41. [19] Athmer CJ, Ho SV, Hughes BM. Results of the Lasagna phase II, Field Experiment. Extended abstract. Emerging technologies in hazardous waste management IX, 15±17, September 1997.