Introduction
Requirements for BTEX Biodegradation
Electron Acceptors and Electron Tower Theory
Evidence of Biological Degradation of BTEX
Compounds
References
There are an estimated 1 to 2 million underground storage tanks containing gasoline in the United States. Of this number it is estimated that 100,000 to 400,000 are leaking either into the soil or directly into the groundwater (Atlas and Cerniglia, 1995). In addition to leaking underground storage tanks, leakage from the subterranean portion of tanks at fuel storage facilities, such as the tank shown below, are contributing to the volume of petroleum hydrocarbons contaminating the subsurface environment. The United States Environmental Protection Agency (EPA) estimates that roughly 11 million gallons of gasoline per year are lost due to leaking underground storage tanks. Gasoline as well as other fuels contain benzene, toluene, ethylbenzene, and the xylene isomers (collectively known as BTEX) which are hazardous compounds regulated by the EPA. These BTEX compounds may comprise greater than 60% of the mass that goes into solution when gasoline is introduced to water (Barbaro et al., 1992). Due to their relatively high solubility, BTEX compounds are the hydrocarbons most frequently reported as groundwater contaminants. Considering that gasoline leaks from underground storage tanks are a major source of groundwater contamination and that almost 50% of the drinking water supply in the United States comes from groundwater wells, there is high potential for drinking water contamination resulting from leaking underground storage tanks.
Petroleum storage tank with portion below grade. Photo by: J.S. Brauner
Once groundwater contamination has occurred, there are numerous techniques available to either contain the pollutant or treat the aquifer. These techniques can range from ex-situ technologies such as excavation and subsequent treatment of aquifer material or pump and treat methods, to physical containment via slurry walls and other impermeable structures, to in situ (a.k.a. intrinsic) remediation via biological and/or chemical transformation of hazardous materials into either less toxic or non-toxic compounds. Of these various techniques, in situ biologically-based remediation has the potential to provide an efficient and cost effective remediation procedure while minimizing site disturbance. Although some recent research has suggested that naturally occurring biodegradation can adequately reduce contaminant concentrations to acceptable levels before the plume reaches a potentially harmful location (thereby reducing the need for engineered treatment methods), no site contaminated with petroleum hydrocarbons has been successfully remediated to drinking water standards solely by in situ bioremediation as of the writing of this document.
It is important to note that bioremediation is not a new technology. In fact, natural microbiological processes have been manipulated for almost a century in wastewater treatment (Baker and Herson, 1994). What is new is the application of this technology to groundwater contaminants in situ. In situ bioremediation refers to the use of natural microbiological processes occurring in the subsurface environment to breakdown complex compounds into simpler, non-toxic compounds without removal of aquifer material. For biodegradation to occur in the subsurface environment, the following basic requirements need to be met (Bedient et al., 1994):
In situ bioremediation can be divided into natural and enhanced methods. As noted in the previous section, naturally occurring bioremediation occurs when a sufficient energy source, carbon source, electron acceptor concentration, and nutrient concentration are available to a native biological population. The rate of naturally occurring bioremediation of BTEX compounds is often limited by either the concentration of an appropriate electron acceptor or a nutrient needed during the transformation of the BTEX into a non-toxic compound. Enhanced in situ bioremediation attempts to stimulate biodegradation by adding either the limiting electron acceptor or the appropriate nutrients to the subsurface environment until the (hydro)carbon substrate becomes the limiting factor in reaction kinetics.
Typical electron acceptors utilized by microorganisms are oxygen, nitrate, iron (III), sulfate, and carbon dioxide. When oxygen is utilized as the electron acceptor, microbial respiration is termed aerobic. When other electron acceptors are utilized, it is termed anaerobic. Depending on the mode of respiration, microbes can be classified into three categories: (1) aerobic; (2) anaerobic; and (3) facultative (Tchobanoglous and Schroeder, 1985). Aerobes thrive only in oxygenated environments using dissolved oxygen as an electron acceptor. Strict anaerobes grow only under highly reduced conditions, where oxygen is effectively absent. Strict anaerobes use electron acceptors such as sulfate or carbon dioxide. Many microorganisms are able to adapt to both aerobic and anaerobic conditions, but are typically more active in the presence of oxygen (Hutchins, 1991). These organisms are termed facultative, and most microbes utilizing nitrate as an electron acceptor tend to be facultative (Firestone, 1982). The following figure adapted from Jorgensen (1989) illustrates the sequence and products of electron acceptor utilization for oxidation of organic carbon.
The electron tower, as depicted above, relates the amount of energy a given microbial population can gain from electron acceptor to the electron acceptors position on the 'tower'. Microbes tend to oxidize organic substrates by the using the electron acceptor that provides the most energy (Stumm and Morgan, 1981). Note that oxygen, which is at the top of the electron 'tower', provides microbes with more free energy (via oxygen reduction) than any other electron acceptor. Carbon dioxide, which is used as the electron acceptor by methanogenic bacteria, yields the least energy of all the electron acceptors, and is therefore located at the bottom of the 'tower'. Thus, the electron tower provide above schematically depicts the order of electron acceptor utilization based on the free energy a given microbial population can gain from reduction of a given electron acceptor.
Field evidence also seems to suggest that natural in situ bioremediation may employ different electron acceptors at various locations throughout a given site. Lyngkilde et al. (1991) report an electron acceptor utilization order determined by measuring field concentrations for each electron acceptor. The trends observed at this observed at this site, as schematically depicted in the figure below, indicate that respiratory conditions of the plume vary from highly reactive aerobic conditions, through anoxic nitrate and iron reduction, to highly reduced sulfate and methanogenic conditions. Note that the order of electron acceptor respiration agrees well with the utilization order predicted by the electron tower theory.
Figure adapted from Lyngkilde et al., 1991.
Source: Brauner (1995)
Early research into in situ bioremediation identified only aerobic degradation of petroleum hydrocarbon groundwater contaminants (Jamison et al., 1975, Raymond et. al., 1976). Further research, however, indicated that petroleum hydrocarbons, such as benzene, toluene, ethylbenzene, and xylenes (the BTEX compounds), are some of the most aerobically biodegradable found in the subsurface environment (Lee et al., 1987). Petroleum aromatic compounds have been shown to degrade by cleavage of the aromatic carbon ring as shown here for benzene:
Figure courtesy of the Department of Chemistry, VPI&SU.
Although natural or artificial recharge may stimulate aerobic biodegradation by reintroducing oxygen to anaerobic regions, the low solubility of oxygen and the rapid reaction rates typical of aerobic environments can severely limit the aerobic biodegradation of petroleum hydrocarbons (Wilson et al., 1986; Zeyer et al., 1986; Barker et al., 1987).
Recent research has recognized the importance of anaerobic degradation of aromatic hydrocarbons, as shown for the BTEX compounds in Table 1 below. Research has shown degradation of aromatic hydrocarbons using nitrate, iron (III), and sulfate as electron acceptors (see Baker and Herson, 1994 for summary). Methanogenic contaminant reduction (i.e., using carbon dioxide as an electron acceptor to form methane gas) has also been demonstrated, but at significantly slower rates than the other degradation processes (Vogel and Grbic-Galic, 1986). The most recent discovery of iron (III) reduction is evidenced by reduced contaminant concentrations and increased levels of aqueous phase iron (II) under anaerobic conditions (Lovley and Lonergan, 1990; Landmeyer et al., 1996). Additional research shows that iron (III) is preferentially used over sulfate (Chapelle and Lovley, 1992), but also shows that iron (III) utilization is inhibited by the presence of oxygen and nitrate. For the BTEX compounds in particular, research has shown toluene, ethylbenzene, and some of the xylenes to degrade anaerobically, with toluene being the most anaerobically degradable. Anaerobic degradation of ethylbenzene and xylene appear to be most significant when they are cometabolized with toluene (Arcangeli and Arvin, 1992). Evidence of anaerobic degradation of benzene, though, has been inconclusive.
| Compound | Conditions | Test | Reference(s) |
|---|---|---|---|
| Benzene | Aerobic Methanogenic, microcosm Aerobic, batch study |
In situ Field simulation Laboratory |
Wilson et. al., 1986b Wilson et. al., 1986c Tabak et al., 1990 |
| Ethylbenzene | Methanogenic, microcosm Anaerobic, continuous flow column Aerobic batch study Nitrate reducing |
Laboratory Field sample Laboratory Laboratory |
Wilson et al., 1986 Kuhn et al., 1988 Tabak et al., 1990 Arcangeli and Arvin, 1994. |
| Toluene | Aerobic, microcosm Aerobic Aerobic Methanogenic, microcosm Anaerobic, continuous flow column Aerobic, batch study Anaerobic, continuous flow column Iron reducing, batch study Aerobic, batch study Methanogenic, microcosm Nitrate reducing microcosm and field study |
Field sample In situ In situ Field sample Field sample Field sample Field sample Laboratory, pure culture Laboratory Field sample Field sample |
Wilson et al., 1983b Wilson et al., 1986a Wilson et al., 1986b Wilson et al., 1986c Zeyer et al., 1986 Swindoll et al., 1988 Kuhn et al., 1988 Lovley and Lonergan, 1990 Tabak et al., 1990 Beller et. al., 1991 Barbaro et al., 1992 |
| m-Xylene | Aerobic Anaerobic, continuous flow column Aerobic, batch study |
In situ Field sample Laboratory |
Wilson et al., 1986b Kuhn et al., 1988 Tabak et al., 1990 |
Anaerobic biodegradation rates for aromatic hydrocarbons are typically an order of magnitude or more less than aerobic rates. However, anaerobic biodegradation may still significantly influence substrate reduction due to longer reaction times (Wilson et al., 1986; Hutchins, 1991; Landmeyer et al., 1996). These longer reaction rates can be attributed to the inefficiency of anaerobes relative to aerobes (Zehnder and Stumm, 1988) and the inhibitory effects of alternate electron acceptors present in the subsurface, especially oxygen (Hutchins, 1991; Chang et al., 1993). The most important factor in determining if a contaminant plume can be successfully remediated may be the identification of the terminal electron accepting process.
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Student Authors: J. Steven Brauner & Marc Killingstad
Faculty Advisor: Daniel Gallagher, dang@vt.edu
Copyright © 1998 Daniel Gallagher
Last Modified: June 7, 1998
