Previous studies examining the biodegradation of organic compounds in fractured rock or karst settings generally have focused on biodegradation in the overburden or regolith above bedrock (Christensen and others, 1997). Wiedemeier and others (1998) described a case study in New Hampshire, where biodegradation of benzene, toluene, and TCE in overburden was documented by using microcosms and field data. Few studies have examined biodegradation in bedrock aquifers. Major and others (1995) determined that microorganisms in a fractured bedrock aquifer in New York dechlorinated TCE to ethene. Acetone and methanol, which were also present in the aquifer, served as electron donors, and SO42- and CO2 were the predominant electron acceptors. Chapelle and others (1995) examined the microbial degradation of organic carbon in the Floridan aquifer, which consists of Tertiary limestone characterized by locally extensive conduits and karst topography. These researchers experienced difficulty in using patterns of electron acceptor consumption as a means to identify discrete oxidation-reduction potential zones in the aquifer; however, they did identify sulfate-reducing and methanogenic conditions in the aquifer.
The lack of studies examining biodegradation in karst aquifers may be due to the widespread perception that contaminants are rapidly flushed out of karst aquifers. In highly developed and well-connected conduit systems, the rate of contaminant migration is expected to be much faster than the rate of biodegradation. Field (1993) states that remediation techniques such as ground-water extraction or bioremediation are impractical in karst aquifers dominated by conduit flow; however, he also states that the belief that contaminants are rapidly flushed out of karst aquifers is a popular misconception. Large volumes of water may be trapped in fractures along bedding planes and other features isolated from active ground-water flow paths in karst aquifers (Wolfe and others, 1997). In areas isolated from the major ground-water flow paths, contaminant migration may possibly be slow enough that biodegradation could reduce contaminant mass if favorable microorganisms, food sources, and geochemical conditions are present.
Some researchers have implied that natural bioremediation in karst or fractured rock is unlikely to occur because of the microbiological characteristics of karst aquifers. Many microbes attach to the aquifer matrix and are not motile (swimmers); therefore, contact between those microbes and dissolved contaminants would be limited by the lack of surface area in fractures and conduits. However, adequate information does not exist to accept or reject the assumption that nonmotile bacteria dominate bacteria communities in karst conduits. Other researchers have implied that biodegradation in bedrock aquifers is limited by a lack of appropriate microorganisms. Vogel (1994), for example, states that hydrologic and geologic characteristics that would not benefit the natural bioremediation of chlorinated solvents include fractured rock systems where small microbial populations exist.
A few reports in the literature dispute that statement. Typical microbial numbers in material from unconsolidated aquifers have been reported to range from 1 x 104 to 1 x 107 cells/mL (Ghiorse and Wilson, 1988; Bedient and others, 1994). Studies have shown that water from bedrock (granite and karst) aquifers also may contain microbial populations within this range. For example, total microbial populations of approximately 1 x 105 cells/mL were detected in ground-water samples from a deep granite formation [800 meters (m) below land surface] in Sweden (Pedersen and Ekendahl, 1990). Total microbial populations of 9.7 x 105 to 8.5 x 106 cells/mL and heterotrophic bacteria populations of 3.5 x 103 to 5.0 x 105 cells/mL were detected in ground-water samples from a gasoline-contaminated karst aquifer in Missouri (O'Connor and Brazos, 1991).
Ground-water studies have detected significant bacteria concentrations in water samples from karst aquifers in Tennessee. For example, fecal coliform and fecal streptococci bacteria were detected and cultured in ground-water samples from wells completed as deep as 90 m below land surface in Williamson County in Middle Tennessee (Hanchar, 1991). Fecal coliform and fecal streptococci bacteria also have been detected in several wells in Bedford and Coffee Counties in Middle Tennessee (Roman-Mas and others, 1991) and in Grainger County in East Tennessee (Weaver and others, 1994). The detection of viable fecal bacteria suggests that a wide variety of other types of bacteria also may be present in karst aquifers in Tennessee.
Microorganisms responsible for chlorinated-ethene degradation have been detected in water samples from bedrock aquifers in other regions. Denitrifying and sulfate-reducing bacteria were identified in nutrient-poor, anoxic ground water from granite formations (240 m below land surface) in Manitoba, Canada, and lab studies indicated that the bacteria were able to mineralize a variety of organic substances (Jain and others, 1997).
Aerobic and anaerobic bacteria have several characteristics that may improve their ability to degrade chlorinated ethenes in karst aquifers. These characteristics include diverse metabolisms and the ability to withstand fluctuating anaerobic and aerobic conditions. These bacteria can derive energy from a wide variety of foods. Research has shown that methanotrophs are capable of using ammonia instead of methane (Bedard and Knowles, 1989; Dalton, 1977) and that ammonia-oxidizing bacteria are capable of using methane (Arciero and others, 1989; Bedard and Knowles, 1989). Research also indicates that ammonia-oxidizing bacteria and methanotrophs have slow death rates and can shift into a dormant stage for extended periods when growth substrates are absent. Methanotrophs also function at low oxygen concentrations and are not inhibited until oxygen is completely consumed (Henry and Grbic-Galic, 1994).
Many bacteria, particularly those adapted for aqueous environments, are capable of moving through the use of flagella. For example, the soil bacteria order Pseudomonales have flagella at one or both ends (Chapelle, 1993) allowing them to swim upgradient or downgradient. Many chemolithotrophic bacteria such as ammonia- and sulfur-oxidizers develop flagella to swim towards or away from chemical stimuli as the need arises. At other times they become nonmotile to conserve energy or metabolize material that flows past them (Atlas, 1987).
Further evidence that bacteria are capable of degrading solvents in a conduit was revealed in a laboratory experiment that examined the aerobic degradation of TCE by methanotrophs in a 30-m by 5.0-cm stainless-steel pipe. TCE contaminated water (approximately 2 mg/L) was pumped into the pipe at a rate of 1 liter per minute, and methanotrophs (32 milligrams dry weight per liter) were added to the pipe water at a rate of 0.1 liter per minute. Results indicated that the TCE was reduced by 88 percent as the contaminated water traveled the length of the pipe at a velocity of 110 meters per hour (m/h) (residence time of 0.91 hour) (U.S. Environmental Protection Agency, 1993).
The ground-water velocity used in the experiment was within the range reported for karst conduits (10 to 500 m/h) by Quinlan (1989). The initial TCE concentration in the flow reactor was also within ranges reported for chlorinated solvent contaminated aquifers; however, the concentration of methanotrophs in the flow reactor may have been higher than concentrations typically found in karst aquifers. Assuming that cell mass is equal to 10-12 grams/cell (Chapelle, 1993), the concentration of methanotrophs in the flow reactor would be approximately 3 x 103 cell/mL. This concentration is within the range normally reported for total bacteria concentrations in water from karst aquifers, but the concentration does not account for species diversity.
| Next | Title Page | Table of Contents | List of Illustrations | List of Tables | Conversion Factors |