Session II-5. Intrinsic Bioremediation of Chlorinated Organics

This session involves a directed reading of Chapter 6 Wiedemeier, beginning on page 241. The chapter begins with a very good introduction to contamination problems associated with chlorinated solvents in the subsurface. It will quickly become evident that biodegradation of chlorinated solvents is significantly more complex than biodegradation of petroleum hydrocarbons, as presented in the previous session. We will emphasize  fundamental concepts as presented in Wiedemeier while, simultaneously, attempting to provide some additional perspective comments. The format is very similar to the previous session. We will identify specific sections in Chapter 6, Wiedemeier which should be read in detail. Comments and questions for these sections will be offered in chronological order.

Introduction, pg 241 Wiedemeier

The most common chlorinated solvents of environmental concern include PCE, TCE, TCA, CT. These solvents will be discussed individually in this chapter. Note that these solvents do not all biodegrade by the same processes. These solvents, as well as many of their degradation products, are highly toxic and/or carcinogenic. Microbial degradation of chlorinated solvents is not as well understood as biodegradation of petroleum hydrocarbons.

Overview of Chlorinated Solvent Biodegradation Processes, pg 242

In the biodegradation process:

Not all solvents are biodegraded by the same processes. See Table 6.1, pg 243.

Aerobic cometabolism is discussed on page 242 and again on page 269. Usually this process must be engineered in the subsurface in order to occur and, therefore, we will not treat it further in this course. Anaerobic cometabolism will likewise not be treated herein.

Aerobic oxidation of Chlorinated solvents is discussed on page 268 and 269. Note that highly chlorinated compounds (PCE, TCE, TCA) are not known to be degraded aerobically, but less chlorinated compounds  (VC, DCE, DCA) can be aerobically degraded if oxygen is present (again refer to Table 6.1, pg 243, Wiedemeier). Aerobic oxidation of lesser chlorinated compounds is attractive at bioremediation sites because it occurs rapidly and usually results in complete mineralization of the contaminant. However, if sufficient oxygen is not present it must be supplied through some type of engineered system.

Anaerobic oxidation of chlorinated compounds is discussed on pg 269. Note that this process is still very much in the research stage and very little is known about it.

Halorespiration of Chlorinated Solvents (Reductive Dechlorination Driven by Hydrogen), pg 243-267

Halorespiration of chlorinated solvents is a form of reductive dechlorination driven by hydrogen as the electron donor. Because Halorespiration is the most important process for intrinsic bioremediation of chlorinated solvents, considerable attention will be given to understanding the basic concepts. The discussion of Halorespiration/reductive dechlorination concepts occurs in Wiedemeier pg 243 – 267. We will proceed by highlighting important concepts presented in Wiedemeier.

Reductive Dechlorination/Halorespiration, pg 243-245. Reductive dechlorination is  a reaction in which a chlorinated solvent acts as an electron acceptor and a chlorine atom is replaced with a hydrogen atom. If this reaction is biologically mediated (i.e. if microorganisms are involved) its termed “Halorespiration”. In halorespiration hydrogen is used  as the electron donor, while the chlorinated compound is used as the electron acceptor.  (see Figures 6.1 and 6.2). Halorespiration probably accounts for the majority of chlorinated solvent natural attenuation at field sites.

Written Assignment: Questions II-17 through II-21.

Conditions for Halorespiration, pg 247-248. Three conditions are necessary for halorespiration to occur in the subsurface: 1) the subsurface environment must be anaerobic and have a low oxidation-reduction potential, 2) chlorinated solvents present must be amenable to halorespiration, and 3) there must be an adequate supply of fermentation substrates for the production of dissolved hydrogen. Halorespiration (reductive dechlorination) will not occur until the subsurface becomes sufficiently reduced to support fermentation. That is, halorespiration will not occur in either aerobic or nitrate reducing environments. Halorespiration will likely occur if sulfate is being consumed (sulfate reduction) or if methane is being produced (methanogenesis). See Figure 4.3 pg176 for redox conditions favorable for halorespiration. Although hydrogen is the most important electron donor in halorespiration there are other fermentation products which may also serve this function. The importance of hydrogen in halorespiration (reductive dechlorination) has only recently been recognized. Relevant research projects conducted prior to this realization are summarized on page 248. Generation of dissolved hydrogen by fermentation is discussed on pg 249. Also recall discussions of fermentation in Chapter 4, pg 169.

Written Assignment: Questions II-22 through II-25.

 

Fermentation of BTEX and Other Substances, pg 250-252. It is very important to understand that petroleum hydrocarbons, including BTEX, can be fermented to produce hydrogen for reductive dechlorination. This means that halorespiration can occur unaided in a region where a chlorinated solvent plume is mixed with a BTEX plume. The fermentation of BTEX is discussed on pg 250. Remember that fermentation is the first step in the oxidation of petroleum hydrocarbons via methanogenesis (pg 205, chapter 5, Wiedemeier). Thus the presence of methane is clear evidence that fermentation is occurring. In summary hydrogen is generated by fermentation of nonchlorinated organics including BTEX, acetone, naturally occurring organic carbon other compounds. Hydrogen is a highly reduced molecule (Figure 4.4 pg 177) and is an excellent electron donor. A wide variety of bacteria can use hydrogen as an electron donor including denitrifiers, iron reducers, sulfate reducers, methanogens, and halorespirators. Thus the production of hydrogen through fermentation does not, by itself, guarantee that hydrogen will be available for halorespiration.

The Monod Kinetic Model, pg 252-254. The Monod kinetic model is useful for describing bacterial growth under substrate (hydrogen) limiting conditions. Monod parameters for halorespirators vs denitrifiers, sulfate reducers  and methanogens are shown in Table 6.3 pg 253. These parameters show that halorespirators will outcompete methanogens and sulfate reducers while denitrifiers will likely outcompete halorespirators. This suggests that high nitrate concentrations in groundwater will very likely prove unfavorable for halorespiration. The likely “chain of events’ leading to the onset of halorespiration is summarized on page 254.

Electron Acceptor/Electron Donor Model for Characterizing Oxidation-Reduction Potential at Chlorinated Sites, pg 255-262. Here is a discussion of an electron donor/electron acceptor model for characterizing oxidation-reduction potential at chlorinated solvent sites. Read this section paying particular attention to Table 6.4, Figure 6.5, and Table 6.5. The  important concept here is that the more oxidized electron acceptors (i.e. nitrate, iron (III) ) are able to extract more energy per mole of hydrogen consumed and, therefore, are able to utilize hydrogen at relatively low concentrations. As can be seen in Figure 6.5,Table 6.5, and Figure 6.6 optimum hydrogen concentrations for halorespirators are on the order of 1 nM. Figure 6.7 shows the flow of donors to acceptors for competing anaerobic reactions. Figure 6.8 shows the thermodynamic flow of electron donor and electron acceptor pathways at chlorinated solvent sites undergoing halorespiration. This figure represents a very good summary of  important concepts discussed previously. It is very important to remember that, in reality, aquifers are heterogeneous and poorly mixed; therefore the oxidation-reduction potential can vary greatly from point to point.

Stoichiometry of Reductive Dechlorination, pg 262-268. The section describing “stoichiometry of reductive dechlorination” serves as a useful reference. The section describing the “microbiology and biochemistry of halorespirators” summarizes research on both pure and mixed cultures of halorespirators. It should be noted that complete transformation of PCE  to ethene has been observed by mixed field cultures while only one pure cultures  has been identified which carries out complete biotransformation. Chlorinated solvents which are amenable to halorespiration are summarized on pg 264. Figures 6.9, 6.10 and 6.11 show the reaction sequences and relative rates for halorespiration of  PCE, TCA and CT.

Written Assignment: Questions II-26, II-27.

 

Oxidation of Chlorinated Solvents, pg 268-269

Material related to oxidation (anaerobic and aerobic) of chlorinated solvents along with cometabolic degradation is summarized on these pages. Additional information on these topics is presented in following sections along with information on bioaugmentation and use of zero-valent iron for treatment of chlorinated plumes.

 

Direct Aerobic Oxidation of Chlorinated Compounds. The following is a summary of material in Wiedemeier supplemented with citations from recent literature.

Both VC and DCE have been shown to degrade via direct aerobic oxidation at rates far in excess of those achieved via halorespiration.  In laboratory-scale experiments using stream-bed sediments containing microorganisms adapted to DCE-containing wastewater, Bradley and Chapelle (1998) measured rapid mineralization of DCE and VC.  Both DCE and VC degradation in these experiments followed Michaelis-Menten kinetics (i.e., first order at low substrate concentrations, zero-order at high substrate concentrations).  The half-saturation coefficients measured in these experiments were 1160 μg/l and 790 μg/l for DCE and VC, respectively, indicating first-order degradation at concentrations relevant to the CBR site.  The first-order rates measured by Bradley and Chapelle (1998) were 91 yr-1 and 365 yr-1 for DCE and VC, respectively.  Comparing these rates with the highest (anaerobic) rates estimated for DCE and VC  indicates increases of 36-fold and 146-fold, respectively, for these compounds.  While it is unlikely that such increases could be realized in a field situation, these figures nonetheless demonstrate that aerobic degradation of DCE and VC may offer an increase in current field rates.  Davis and Carpenter (1990) also noted rapid degradation of VC under aerobic conditions in soil-groundwater microcosm experiments.  Although the observed first-order decay rate was only 4.4 yr-1, the microbial consortium used had no previous exposure to chlorinated solvents.  Obviously, generation of aerobic conditions at many field sites would require the subsurface introduction of oxygen.  Methods for accomplishing this include direct injection of air or oxygen into groundwater, or the addition of a solid phase media which releases oxygen as a dissolution process (e.g., Oxygen Release Compound, Regenesis, Inc.). 

References cited:

Bradley, P.M. and F.H. Chapelle, 1998, Effect of contaminant concentration on aerobic microbial mineralization of DCE and VC in stream-bed sediments, Environ. Sci. Technol., vol. 32, pp. 553-557. 

Davis, J.W., and C.L. Carpenter, 1990, Aerobic biodegradation of vinyl chloride in groundwater samples.  Appl. Environ. Microbiol., vol. 56, pp 3878.

Aerobic Cometabolism of Chlorinated Compounds. The following is a summary of material in Wiedemeier supplemented with citations from recent literature.

With the exception of PCE, all chlorinated ethenes and ethanes have been shown to undergo cometabolic degradation under aerobic conditions (Murray and Richardson, 1993; McCarty and Semprini, 1994), although biotransformation rate data is not available in the literature.  Aerobic cometabolism is accomplished when aerobic bacteria produce enzymes such as oxygenase  to degrade a primary metabolite such as methane (Bouwer, 1994).  These enzymes then fortuitously degrade the chlorinated solvent, with no benefit to the bacteria (Lee, et al., 1998).  Several compounds, namely TCA and DCE have been shown to be difficult to degrade under aerobic cometabolism.  As with aerobic oxidation, more highly chlorinated compound is, the lower the rate of cometabolic degradation (McCarty and Semprini, 1994).  Aerobic cometabolism requires the presence of a suitable primary electron donor (i.e toluene or methane) in addition to oxygen.  Like direct aerobic oxidation, aerobic cometabolism would require the addition of oxygen to the subsurface. 

References Cited:

Bouwer, E.J., 1994, Bioremediation of chlorinated solvents using alternative electron acceptors, in Handbook of Bioremediation, R.D. Norris et al., eds., Lewis Publishers, Boca Raton, FL.

Lee, M.D., J.M. Odom, and R.J. Buchanan, Jr., 1998, New perspectives on microbial dehalogenation of chlorinated solvents: insights from the field, Annu. Rev. Microbiol., vol. 52, pp. 423-452.

McCarty, P.L. and L. Semprini, 1994, Groundwater treatment for chlorinated solvents, in Handbook of Bioremediation, R.D. Norris et al., eds., Lewis Publishers, Boca Raton, FL.

Murray, W.D. and M. Richardson, 1993, Progress toward the biological treatment of C1 and C2 halogenated hydrocarbons, Crit. Rev. Environ. Sci. Technol., vol. 23, pp. 195-217.

Direct Anaerobic Oxidation of Chlorinated Solvents. The following is a summary of material in Wiedemeier supplemented with citations from recent literature.

Anaerobic oxidation occurs when bacteria utilize the chlorinated organic as the electron donor and some compound other than oxygen as the electron acceptor.  Highly oxidized chlorinated solvents such as PCE, TCE, and TCA are not oxidized under anaerobic conditions because this reaction is energetically unfavorable (Bouwer, 1994).  Instead, these tri- and tetra-chlorinated compounds are degraded through anaerobic halorespiration, where the chlorinated compound acts as the electron acceptor rather than the electron donor.  Mono- and di-chlorinated compounds (VC, DCA, DCE), however, can be biodegraded through anaerobic oxidation.  For example, anaerobic oxidation of VC to CO2 has been demonstrated under manganese-reducing, iron-reducing, and sulfate-reducing conditions in laboratory systems (Bradley and Chapelle, 2000).  There is further evidence to indicate that the more energetic anaerobic electron acceptors, such as nitrate and Fe(III), are capable of promoting anaerobic degradation of some solvent compounds.  Under denitrifying conditions at a field site, Thomson et al. (1995), measured first order decay constants of 6.3 yr-1 for 1, 2-DCA and 4.2 yr-1 for both 1,1-DCA and VC.  In this work, nitrate was injected into groundwater to achieve a final concentration of 50 mg/l (as N). 

Anaerobic biodegradation of DCE and VC has also been reported under conditions of Fe(III) reduction in laboratory microcosms using sediments from a solvent contaminated field site (Bradley and Chapelle, 1996; Bradley and Chapelle, 1997).  Under Fe(III) reducing conditions, VC and DCE degradation was found to be first-order in the range of concentrations found at the CBR site, with rate constants of 91 yr-1 and 219 yr-1, respectively (Bradley and Chapelle, 1997).  In the case of these experiments, Fe(III) was supplied in the form of Fe-EDTA.  The initial concentration of iron was approximately 500 mg/l as Fe(III) and measurement of reduced iron at day 37 of the study indicated the production of 11 mg/l Fe(II), suggesting iron reduction as the mechanism of biotransformation.  Although Fe(III) reduction has been shown to be possible in laboratory systems, reports of successful field application are lacking. 

References Cited:

Bouwer, E.J., 1994, Bioremediation of chlorinated solvents using alternative electron acceptors, in Handbook of Bioremediation, R.D. Norris et al., eds., Lewis Publishers, Boca Raton, FL.

Bradley, P.M. and F.H. Chapelle, 1996, Anaerobic mineralization of vinyl chloride in Fe(III)-reducing aquifer sediments, Environ. Sci. Technol., vol. 30, pp. 2084-2086.

Bradley, P.M. and F.H. Chapelle, 1997, Kinetics of DCE and VC mineralization under methanogenic and Fe(III)-reducing conditions, Environ. Sci. Technol., vol. 31, pp. 2692-2696.

Bradley, P.M. and F.H. Chapelle, 2000, Acetogenic microbial degradation of vinyl chloride, Environ. Sci. Technol., vol. 34, pp. 2761-2763. 

Thomson, J.A.M., M.J Day, R.L. Sloan and M.L. Collins, 1995, In situ aquifer bioremediation at the French Limited Superfund Site, in Applied Bioremediation of Petroleum Hydrocarbons, eds., R.E. Hinchee, J.A. Kittel, and H.J. Reisinger, pp. 453-459, Battelle Press, Columbus, OH.

Bioaugmentation to Facilitate Chlorinated Solvent Bioremediation. Review of current literature.

Bioaugmentation is the addition of a selected consortium (or single species) of microorganisms to assist biotransformation in situ.  The added organisms can be isolated and enriched from the site in question or from another site.  While typically considered unnecessary for petroleum hydrocarbons, bioaugmentation has been shown to be effective under some conditions at solvent-impacted sites.  It has been noted at many anaerobic solvent sites that DCE is often present at greater concentrations than parent ethene compounds.  This may be due to either slower degradation of DCE via halorespiration or the prevalence of microorganisms that reduce PCE and TCE to DCE, but no further (Wiedemeier et al., 1999).  Support for latter explanation suggests that a single microbial species (or several closely related species) may be responsible for complete dechlorination of more highly chlorinated compounds.  To date, only Dehalococcoides ethenogenes strain 195 has been found to be capable of complete dechlorination of the range of chloroethenes (Maymó-Gatell, et al., 1999).  Although this organism was originally isolated from municipal sewage sludge, it has recently been isolated from diverse and widespread geologic and industrial sites.  Hendrickson et al. (2002) sampled 24 field sites in the USA and Europe contaminated with PCE, TCE, DCE, and VC, and found D. ethenogenes strain 195 or a closely related organism as identified by 16S rDNA sequencing) to be present at 21 of the sites.  Interestingly, only sites without an active population of D. ethenogenes were observed to have significant accumulations of DCE.  At all other sites, dechlorination proceeded to ethene without significant buildup of DCE or VC. 

Bioaugmentation has been successfully implemented at a TCE-contaminated aquifer at Dover Air Force Base, DE (Ellis et al., 2000).  Halorespiration at this site did not proceed past DCE prior to subsurface injection of a microbial enrichment derived from a solvent contaminated site in Largo, FL.  Following enrichment and injection of this culture, TCE and DCE were fully converted to ethene, and the injected culture was found to persist at the site.  While D. ethenogenes was found to be present in the injected inoculum, 16SrRNA analysis indicates that it did not constitute a major part of the consortium.  In this case, it is likely that D. ethenogenes and other organisms capable of halorespiration grow interdependently. 

References Cited:

Ellis, D.E., D.J. Lutz, J.M. Odom, M.D. Lee, M.R. Harkness, and K.A. Deweerd, 2000, Bioaugmentation for accelerated in situ anaerobic bioremediation, Environ. Sci. Technol., vol. 34, pp. 2254-2260.

Maymó-Gatell, X., V. Tandoi, J.M. Gossett, and S.H. Zinder, 1999, Reductive dechlorination of chlorinated ethenes and 1,2-Dichloroethane by “Dehalococcoides ethenogenes” 195. Appl. Environ. Microbiol., vol. 65, pp. 3108-3113.  

 

Zero-Valent Iron for Abiotic Degradation of Chlorinated Solvents. Review of Current Literature.
                                                                                               
Use of zero-valent iron [Fe(0)] as a means of dechlorinating ethenes and ethanes has recently emerged as a viable alternative for field-scale remediation.   Fe(0) is typically employed in a permeable reactive barrier (PRB) which is placed in the subsurface saturated zone perpendicular to groundwater flow.  PRBs using Fe(0) promote chlorinated solvent biotransformation both because they act to reduce the oxidation-reduction potential of groundwater (making halorespiration more favorable) and because they produce an abundance of electrons which react abiotically to dehaologenate chlorinated solvents (Roberts et al., 1996).  At the CBR, subsurface conditions are already conducive to halorespiration, so the main benefit of a Fe(0) barrier would be abiotic reaction.  Dehalogenation of chlorinated ethenes via Fe(0) proceeds via hydrogenolysis (replacement of a chlorine by a hydrogen), and sequentially reduces CE→TCE→ DCE→VC, where each step in the process consumes 2 electrons and one H+ (Roberts et al., 1996).  Consequently, the process requires a ready supply of electrons (which are supplied by the zero valent iron) and consumes acidity, typically raising the pH in the vicinity of the PRB to the range of 9-10.  Under laboratory conditions, rapid decay of chlorinated ethenes has been reported using Fe(0).  Gillham and O’Hannesin (1994) measured first order decay constants of 446 yr-1, 14 yr-1, and 16 yr-1 for TCE, cis 1,2-DCE, and VC, respectively.  These rates for DCE and VC decay are approximately one order of magnitude greater than currently observed at the site. 

The rate of dechlorination resulting from zero-valent iron has been found to be highly correlated with the surface area of Fe(0) available for reaction (Gillham and O’Hannesin, 1994) and the condition of the iron surface with regard to corrosion (Farrell et al., 2000).  Corrosive decay of the iron surface resulted in a 6-fold decrease in TCE decay rates after approximately 2 years of operation (Farrell et al., 2000) in column tests.  A zero-valent iron field scale barrier was installed in Denver in 1996 to treat a complex mixture of chlorinated ethenes and ethanes.  TCA, DCA, TCE, 1,1-DCE and VC were measured upstream from the PRB at concentrations of 200, 15, 600, 230, and 18 μg/l, respectively. Downstream from the PRB, only DCA was measured at a concentration of 15 μg/l (McMahon et al., 1999).  First-order degradation rates for TCA, TCE, and DCE were measured at 280 yr-1, 245 yr-1, and 114 yr-1, respectively.  In this field implementation, pH increased from 7.1 upstream from the barrier to 9.7 within the barrier. 

References Cited:

Gillham, R.W., and S.F. O’Hannesin, 1994, Enhanced degradation of halogenated aliphatics by zero-valent iron, Ground Water, vol. 32, pp. 958-967.

Farrell, J., M. Kason, N. Melitas, and T. Li, 2000, Investigation of the long-term performance of zero-valent iron for reductive dechlorination of trichloroethylene, Environ. Sci. Technol., vol. 34, pp. 514-521.

McMahon, P.B., K.F. Dennehy, and M.W. Sandstrom, 1999, Hydraulic and geochemical performance of a permeable reactive barrier containing zero-valent iron, Denver Federal Center, Ground Water, vol. 37, pp. 396-404.

Roberts, A.L., L.A. Totten, W.A. Arnold, D.R. Burris, and T.J. Campbell, 1996, Reductive elimination of chlorinated ethylenes by zero-valent metals, Environ. Sci. Technol., vol. 30, pp. 2654-2659.

 

Classification System for Chlorinated Solvent Plumes, pg 270-277

A proposed classification system for chlorinated solvent plumes in the field is presented ( pg 270-277) based on the amount and origin of fermentation substrates that produce hydrogen which drives halorespiration. Subsurface environments are classified into three types: Type 1, systems that are anaerobic due to anthropogenic carbon, Type 2, Systems that are anaerobic due to naturally occurring carbon, and Type 3, Aerobic systems due to the absence of fermentation products. This material provides valuable insight for interpreting the size, shape and migration of chlorinated solvent plumes under a variety of field conditions. It also illustrates how many of the fundamental concepts regarding halorespiration can be applied to interpret results from field studies or make predictions of future solvent plume behavior.

Chlorinated Solvent Rate Data from the Literature, pg 278

This sections summaries what is know regarding biodegradation kinetics of chlorinated solvents. As can be seen from Table 6.6  observations of rate coefficients and half-lives (both laboratory and field) vary considerably. Table 6.7 provides recommended guidelines for choosing first order rate coefficients or half lives for selected solvents.

Chlorinated Solvent Plume Data Bases, pg 281

Tables 6.9 – 6.11 provide characteristics and data for a large number of chlorinated solvent plumes observed under various field conditions

Written Assignment: Question II-28.