Bioremediation is a term that refers to a number of remediation technologies for treatment of both soil and groundwater using microorganisms (USEPA, 2012). Bioremediation is typically used to treat sites contaminated with organic substances (USEPA, 2001a), but it can also be used to immobilize inorganic contaminants such as heavy metals, although this is a developing area (Sharma & Reddy, 2004). Of sites where bioremediation is used 21% are considered petroleum sites and 31% are wood preserving sites, 19% are landfill sites and 13% are underground storage tanks (USEPA, 2001a). The remaining 16% of sites are less common and include fire training areas, vehicle maintainance areas, pesticide manufacturing sites and spills (USEPA, 2001a). The most common organic contaminants typically include polycyclic aromatic hydrocarbons (PAHs) (e.g. benzene, toluene, ethylbenze and xylenes (BTEX)), polychlorinated biphenyls (PCBs), pesticides and herbicides (Sharma & Reddy, 2004), (USEPA, 2001a).
Fundamentally, bioremediation uses microbes (e.g. bacteria, yeast, and fungi) to ‘digest’ toxic organic contaminants (Sharma & Reddy, 2004), such as oil, producing non-toxic products such as water and carbon dioxide (USEPA, 2001b). The process of breaking down organic contaminants with microorganisms is referred to as biodegradation. This can occur in the presence of oxygen or without oxygen, known as aerobic and anaerobic conditions, respectively. This is shown diagrammatically in Figure 1 and using a simplified equation in Equation 1 below.
Figure 1 – Diagram representing basic concept of bioremediation (USEPA, 2001b)
Equation 1 – Simplified chemical equation of aerobic bioremediation (Sharma & Reddy, 2004)
This is simplification and the actual process is much more complex. Sometimes the ‘native’ population of microorganisms, known as indigenous microorganisms, may not be able to biodegrade the contaminant in the soil. In this case ‘foreign’ exogenous microorganisms must be introduced to the soil (Sharma & Reddy, 2004). The addition of exogenous microorganisms to the soil is known as bioaugmentation. For the microorganisms to survive in the ground, a delicate balance of a number of parameters including temperature, pH, moisture content, oxygen concentration and nutrients in the soil must be achieved (Sharma & Reddy, 2004) and (USEPA, 2012). Bioremediation is most efficient (Sharma & Reddy, 2004):
If this balance is not present in the soil it can be achieved by adding amendments to the soil in a process known as biostimulation (USEPA, 2012). Amendments such as H2O2 (hydrogen peroxide) or Oxygen-Releasing-Compounds (ORCs) are typically pumped or injected into the ground. Figure 2 shows a schematic of the balance of nutrients for microorganisms.
Figure 2 – Schematic of microorganism requirements (USEPA, 2012)
If the adequate conditions in the soil cannot be achieved easily, then ‘in-ground’ or in-situ treatment may not be possible. Instead the soil may be excavated and treated ‘above-ground’ or ex-situ (USEPA, 2012). Typically groundwater contamination is treated in-situ. There are many different treatment methods available and these include (USEPA 2001a) (Sharma & Reddy, 2004).
Bioventing – Introduction of oxygen into the soil using forced air to encourage microbial activity. Nutrients such as nitrogen and phosphorus may also be added to the soil to increase the growth rate of microorganisms.
Injection or pumping of ORCs – ORCs are introduced into the soil to encourage microbial activity (similar to bioventing). This method is only used when the groundwater is already contaminated because ORCs are in liquid form.
Slurry-phase lagoon aeration – Air and soil are combined in a lagoon to encourage microbial activity.
3 varieties of solid-phase treatment:
Figure 3 – Schematic of composting method (USEPA, 2001a)
An aqueous mix of soil and water is mixed with amendments and stored in tanks. The mix is continuously stirred to keep solids suspended to maximize contact area between contaminated soil and microorganisms.
Biosparging – Air is injected into the groundwater to encourage aerobic biodegradation.
Aerobic treatment – Amendments introduced by injection or water recirculating systems into groundwater. Can include ORC or H2O2. Organic contaminants are converted to carbon dioxide or water. Aerobic treatment is faster than anaerobic, and therefore preferred.
Anaerobic treatment – Carbon sources such as molasses or Hydrogen Releasing Compounds (HRC) are introduced by injection or water recirculating systems. In the case of carbon sources and HRCs the contaminants are converted into methane, limited amounts of carbon dioxide, and traces of hydrogen.
A typical groundwater treatment setup is shown in Figure 4.
Figure 4 – Diagram of in-situ groundwater treatment (USEPA, 2001a)
Ex-situ treatment is decreasing in popularity compared to in-situ treatment. In 1991, 35% of treatment work at Superfund sites was in-situ compared to 53% in 1999 (USEPA, 2001a). This is likely to be due to increasing effectiveness of in-situ methods and a desire to avoid exposing the contaminated soil at the surface where likelihood of human contamination is increased. The most popular in-situ and ex-situ treatment methods are bioventing and land farming, respectively.
An interesting use of bioremediation is the treatment of acid mine drainage (AMD) using passive wetlands and sulfate-reducing bacteria. This will be discussed later.
In aerobic degradation successful bioremediation can be monitored by measuring carbon dioxide and oxygen levels. Decreases in oxygen concentration and increases in carbon dioxide concentration signify increased bioactivity.
Bioremediation can be complemented with other technologies such as soil vapor extraction and soil washing.
Acid Mine Drainage (AMD) is a common problem around the world where mining plays an important role in the economy. The passive treatment of AMD using sulfate-reducing bacteria (SRB) in anaerobic wetlands in a common technology employed. It is popular because it requires little maintenance and any contaminants that precipitate out remain in the wetland waters (Johnson & Hallberg, Acid mine drainage remediation options: a review, 2005a). A downside is that it requires significant land space and so can be an expensive option. Anaerobic wetlands can remove many metals including iron, zinc and aluminium from waters draining from mines. Figure 5 shows a typical anaerobic wetland system.
Figure 5 – Typical anaerobic wetland
Biodegradation is an oxidation-reduction reaction involving organic compounds and facilitated by microorganisms (Sharma & Reddy, 2004). It can also be thought of as a process where an organic compound becomes smaller through chemical or biological processes (Fetter, 1993) for example a halide ion may be replaced with a hydrogen ion therefore producing a compound of lower molecular weight. Biodegradation can also be thought of as the metabolism of organic contaminants by microorganisms, i.e. microorganisms break down organic compounds to produce energy and carbon to permit growth (Sharma & Reddy, 2004). Biodegradation can occur in the presence of oxygen and is termed aerobic or without oxygen and is termed anaerobic. In aerobic biodegradation, organic compounds are oxidized by an electron acceptor, which is reduced. If the oxidation of an organic contaminant occurs with the reduction of molecular oxygen it is known as aerobic heterotrophic respiration (Sharma & Reddy, 2004). Anaerobic biodegradation processes include (Sharma & Reddy, 2004):
Bacteria can be classified depending on their functioning in the environment or by how they create energy (Sharma & Reddy, 2004). Classification by function has two types:
Classification by energy creation has three types:
There are 5 main requirements for successful biodegradation to occur (Sharma & Reddy, 2004):
Biodegradation of hydrocarbons, such as alkanes, can occur under aerobic conditions. Bacteria that are able to do this include Micrococcus, Pseudomonas, Mycobacterium and Nocardia (Fetter, 1993). For example the oxidation of n-hexane is as follows (Fetter, 1993):
Branched alkanes are harder for microorganisms to biodegrade then straight chain alkanes (Fetter, 1993).
Halogenated organic compounds can be oxidized or reduced by microorganisms. An example of an oxidation reaction is α-hydroxylation, here a hydrogen atom on a carbon atom that is also bonded to a halogen is replaced by an OH- group (Fetter, 1993). An example is the α-hydroxylation of 1,1-dichloroethane to 1,1-dichloroethanol:
The halogenated oxygen can loose another hydrogen to form an aldehyde:
Reduction reactions can occur in two different processes hydrogenolysis and dihaloelimination (Fetter, 1993). In hyrdogenolysis a halide ion is replaced by a reduced species producing an alkyl radical. The alkyl radical reacts with a H+ ion and is substituted for the negatively charged halide e.g.:
Dihaloelimination occurs when halides are present on adjacent carbon atoms. The halogen removed from each carbon atom creates a double bond between the carbon atoms thus forming an alkene (Fetter, 1993). Dihaloelimination takes a halogenated alkane and produces a halogenated alkene eg. Hexachloroethane to tetrachloroethene:
The aerobic biodegradation of organic compounds, consumption of oxygen and growth of microorganisms can be described by the Michaelis-Menten functions, also known as the Monod Function (Fetter, 1993) and (Sharma & Reddy, 2004):
H – hydrocarbon concentration in the pore fluid (ML-3)
O – oxygen concentration in the pore fluid (ML-3)
Mt – total aerobic microbial concentration (ML-3)
hu – maximum hydrocarbon utilization rate per unit mass of aerobic microorganisms (T-1)
Y – microbial yield coefficient (g cells/g hydrocarbon)
Kh – hydrocarbon half saturation constant for aerobic decay (ML-3)
Ko – oxygen half saturation constant (ML-3)
kc – first order decay rate of natural organic carbon
Coc – natural organic carbon concentration (ML-3)
b – microbial decay rate (T-1)
G – ratio of oxygen to hydrocarbon consumed
Anaerobic biodegradation can be described by a different modified Monod function (Sharma & Reddy, 2004):
Ma – total anaerobic microbial concentration (ML-3)
hua – maximum hydrocarbon utilization rate per unit mass of anaerobic microorganisms (T-1)
Ka – hydrocarbon half saturation constant for anaerobic decay (ML-3)
One important note is that for a given organic substrate, the concentration of that substrate cannot be reduced below the concentration required for the survival of the microorganism. Thus there is always a minimum concentration of organic substrate, Hmin, given by (Fetter, 1993):
When a mix of hydrocarbons is present it may be possible to reduce the concentration below Hmin for a given hydrocarbon if it was the sole contaminant by utilizing other hydrocarbons to allow the microorganisms to survive.
The biodegradability of a contaminant can be determined by carrying out tests to determine the Biological Oxygen Demand (BOD) and Chemical Oxygen Demand. The BOD is a measure of oxygen required by microbes to breakdown a given organic contaminant in a sample of water. The COD is an indirect measure of the amount of organic contaminant present in a sample of water. A ratio of BOD and COD gives an indication of the biodegradability of a given contaminant.
Acid mine drainage (AMD) is a common problem worldwide, an estimated 19,300 km of rivers and 72,000 ha of lakes have been damaged by AMD (Johnson & Hallberg, 2005a). Because of the association with mining and underground activity the effluent of mines is often rich in heavy metals and can contain other contaminants such as arsenic. The main source of AMD is iron pyrite (FeS2). Many metals that are mined occur naturally as metal sulfides and pyrite tends to be found in the same location as the metal ore of interest (Johnson & Hallberg, 2005a). As mining progresses and the water table is pumped down to allow advancement of tunnels and the pyrite is exposed to air and water and sets in motion a series of acid generating reactions. The processes are complicated and varied, but can be summarized as follows (Blodau, 2006):
First pyrite is oxidized releasing Fe2+ ions:
Fe2+ ions are then further oxidized to Fe3+:
The Fe3+ can then go on to oxidize more pyrite in a highly acid generating reaction:
pH is given by:
So clearly, as the number of H+ ions increase the pH will fall becoming increasingly acidic. The high acidity of drainage from mines caused by the above reactions, along with the heavy metals associated with mining pose a significant risk to the environment.
There are many possible methods that can be employed to treat AMD and each has strengths and weaknesses depending on the unique chemistry of each AMD. One popular method for the treatment of net acidic AMD is passive anaerobic wetlands that make use of sulfate reducing bacteria (SRB). A passive treatment method here is defined as a method that makes use of naturally occurring energy and decontaminating agents.
Remediation by SRB can be represented by two chemical equations described by Johnson & Hallberg, (2005a) and Doshi (2006). First sulfate reduction, here organic matter is shown as CH2O:
Then metals may precipitate out as metal sulfides, where Me represents a suitable metal:
Metals that can precipitate out this way include cadmium, copper, iron, lead, mercury, nickel, zinc and manganese. Arsenic, antimony and molybdenum can also form sulfide complexes (Doshi, 2006).
SRB that are suitable to use in anaerobic wetlands include Desulfovibrio, Desulfomicrobium, Desulfobulbus, Desulfobacter, Desulfobacterium, Desulfococcus, Desulfosarcina, Desulfomonile, Desulfonema, Desulfobotulus, and Desulfoarculus (Doshi, 2006). To optimize the efficiency of these bacteria, a balance of temperature and pH must be achieved. Anoxic conditions are also necessary, the introduction of fresh manure helps to promote bioactivity that uses up oxygen present in the water. SRB have been shown to function at pH as low as 2.5 and temperatures as low as 6°C, although optimal temperature is 20°C-40°C (Doshi, 2006).
Like all remediation methods, bioremediation is best suited for certain types of contamination and certain types of soils.
Bioremediation is excellent for biodegrading organic contaminants. Common types of organic contaminant that respond well to bioremediation include: petroleum hydrocarbons, non-chlorinated chemicals (e.g. acetone), wood treating chemicals (e.g. creosote and pentachlorophenol (PCP)), certain chlorinated aromatic compounds (e.g. chlorobenzenes and biphenyls with five or fewer chlorine atoms per molecule) and certain chlorinated aliphatic compounds (e.g. tricholoroethene (TCE)) (USEPA, 2001a).
Bioremediation methods can be implemented to tackle organic contaminants in three key ways (Sharma & Reddy, 2004):
Here destruction of organic contaminants is taken to mean the physiological process by which microorganisms bond with contaminants through oxidation and reduction reactions, the energy released in these reactions provide energy for the microorganism (Sharma & Reddy, 2004). Organic contaminants such as hydrocarbons acts as electron donors in these reactions (aerobic), whereas other common contaminants such as chlorinated solvents can act an electron acceptors (anaerobic).
When applied to waste the suitability of biodegredation techniques for particular types of waste can be assessed by following relevant ASTM standards. For example the biodegradation of plastics in soil and in bioreactor landfills can determined by following ASTM D5988 (Standard Test Method for Determining Aerobic Biodegredation of Plastic Materials in Soil) and ASTM D7474 (Standard Test Method for Determining Aerobic and Anaerobic Biodegredation of Plastic Materials under Accelerated Bioreactor Landfill conditions), respectively.
Inorganic contaminants such as heavy metals from AMD are often treated using bioremediation techniques. Bioremediation can in some cases change the valence state of heavy metals and render it immobile e.g. mobile hexavalent chromium to immobile trivalent chromium.
Almost all contaminated soils with moisture content able to support microbial life are suitable for treatment for using bioremediation. However, soils of low permeability can be difficult when trying to get amendments to permeate throughout the contaminated soil mass (Sharma & Reddy, 2004). If the contaminant concentration is very high, the conditions may be too toxic for microorganisms to survive making bioremediation ineffective.
Homogenous soils with hydraulic conductivity, k, equal to 10-4 cm/s or greater are suitable for treatment with bioremediation (Sharma & Reddy, 2004). If the hydraulic conductivity is too low adding amendments and removing products from the soil may be difficult. Bioremediation can be used to treat Dense Non-Aqueous Phase Liquids (DNAPLs). It is also ideal for treating single sources and as a follow up treatment after free-product removal by a different method.
Advantages of bioremediation include (Sharma & Reddy, 2004) and (Vivaldi, 2001):
Disadvantages of bioremediation include (Sharma & Reddy, 2004):
Necessary equipment includes:
Figure 6 - In-situ bioremediation (Tlusty, 1999)
In-situ bioremediation is aided by biostimulation, enhancing soil with oxygen, moisture, and nutrients, for effective treatment. These constituents are added at injection points, as depicted in the figure.
Figure 7 - In-situ bioremediation – air injection system (Biotechpedia, 2011)
This system helps to remove volatile contaminants from the effected soil. The air injected into the contaminated soil disperses the groundwater, which provides necessary nutrients, such as oxygen, nitrogen, phosphorous, hydrogen, and carbon, for successful bioremediation.
Figure 8- In-situ bioremediation – water circulation system (Biotechpedia, 2011)
This system circulates water through the contaminated water and soil. The water contains nutrients such as oxygen, nitrogen, phosphorous, hydrogen, and carbon that are necessary for successful bioremediation.
Necessary equipment includes:
Figure 9 - Schematic of slurry phase bioremediation (USEPA, 1995a)
In slurry phase bioremediation, contaminated soil is mixed with water to create a slurry, and then aerated. The advantage of this type of bioremediation is that conditions such as pH, temperature, and nutrients are monitored and can be adjusted to aid treatment.
Land farming is a form of solid-phase bioremediation. The process involves spreading the contaminated soil in fields or treatment beds that are an inch thick. The soil is then tilled to allow oxygen into the soil. Treatment is achieved through biodegradation, aeration, and photoxidation. Figure 10 shows an image of land farming in process
Figure 10 - Picture of bioremediation of contaminated soil – landfarming (ETec, 2013)
Windrow systems are a type of composting, another form of ex-situ solid phase bioremediation. Compost is stacked in elongated piles and aeration of the soil is accomplished by tearing down and then rebuilding the piles. Figure 11 shows an example of a windrow
Figure 11 - Ex-situ remediation – picture of windrow (Proper, 2013)
Bioremediation of soils is typically a very affordable technique especially if the main contaminant is organic. Costs range from $30-$750 per cubic yard of treated soil (Sharma & Reddy, 2004). From 1989-1997, the cost of 22 bioremediation projects carried out at superfund sites ranged from $13-$500 per cubic yard of treated soil (USEPA, 2001a). 45 bioventing projects also carried out at superfund sites showed a wide range in costs from $1.36-$333 per cubic yard of soil.
Bioremediation of groundwater is also cheap if the main contaminant is organic. Costs can range form $33-$200 per 1000 gallons of treated water (Sharma & Reddy, 2004). The main cost in groundwater treatment methods usually comes from implementing pumping or injection systems.
For soil treatment, there are a number of considerations. Air emissions must be controlled and seepage of contaminants into the groundwater must be controlled. If ex-situ methods are used then VOCs cannot be allowed to enter the atmosphere (Sharma & Reddy, 2004) unless carefully monitored and released within hazardous air pollution limits.
In the treatment of groundwater, there are other considerations. If water is drawn from the ground it may be considered a contaminated material according to the Resource Conservation and Recovery Act (RCRA) (Sharma & Reddy, 2004). Drawing water may be considered an active management and disposal restrictions may apply. Once contaminant levels are reduced to specific thresholds then the treated water may be injected into a usable aquifer.
Removal of various types of organic contaminants has been successful and unsuccessful at a number of sites (USEPA, 2001a). This is simply a list showing examples of sites where various methods have been either successful or unsuccessful, further details can be found in "Use of Bioremediation at Superfund Sites" (USEPA, 2001a).
Burlington Northern Superfund Site, Minnesota. PAHs were reduced from 70,633 mg/kg to 800mg/kg. A reduction of around 99%. The cleanup goal was 8,632 mg/kg. The contaminated material consisted on soil and sludge and was treated using land farming techniques. The additives used included lime and cattle manure. Bonneville Power Administration Ross Complex. Removal was at 90% but targets were not achieved. The initial PAH concentration was around 150mg/kg, this was reduced to 6.76-21.83 mg/kg. The cleanup goal was 1 mg/kg. In this case only soil was treated using land treatment and UV oxidation as part of a technology demonstration. Peroxide and ethanol were used as amendments.
Groundwater at the Avco Lycoming site was contaminated with TCE at 67μg/L and was reduced to 6.7μg/L. The required level was actually 5μg/L and so the project was considered unsuccessful. The method employed was direct injection under anaerobic conditions. The amendment was molasses. The project was actually a pilot study and one of the first to establish clean up goals (USEPA, 2001a). It was probably unsuccessful because the molasses was unable to mix fully with the contaminated water. Another project at the US Department of Energy (DOE) site at Savannah River, South Carolina was successful. Groundwater contamined with TCE (10-1,031μg/L) and PCE (3-124μg/L) was treated using a recirculating cell with amendments of nitrogen, phosporus and methane. Concentrations of both contaminants was reduced to μg/L as required. Success is attributed to the fact that the recirculating cell ensured a good mix of contaminated water and amendments.
Navajo Indian Reservation. Toxaphene was at 4000mg/kg and reduced to 180mg/kg using anaerobic slurry-phase bioremediation. Removal being at 95%. At the Stauffer Chemical Company site, a cocktail of 6 chemicals including DDT (88.4 mg/kg) was present. A composting method known as Xenorem™ was used as part of a demonstration and DDT concentrations were lowered to 8.91 mg/kg. Target removal levels not achieved were not achieved at the end of the 64 day trial period, however. It has been suggested that a longer test period may have reduced concentrations futher.
Removal of explosives can vary greatly. The Joliet Army Ammunition Plant was contaminated with trinitrotoluene (TNT) and tetryl. Two pilot tests achieved removal of removal of TNT at 31% and 97% and tetryl at 3% and 100%, respectively (USEPA, 2001a). The 'unsuccessful' approach mixed compacted soil with potato waste. Then a blend of aerobic and anaerobic bacteria were added and then nutrients were sprayed onto the soil once every two weeks. The 'successful' approach made use of powdered iron and a proprietary organic amendment (DARAMEND®) mixed with the soil. The amendment alters the physical and chemical properties of the contaminated soil to support microbial life and promote biodegradation. Throughout the treatment conditions cycled between aerobic and anaerobic conditions. From these and other similar projects it has been suggested that the addition of proprietary amendments such as DARAMEND® lead to greater removal of explosive contaminants.
French Limited in Crosby, Harris County, Texas (EPA Region 6) was a 25-acre sand mining site from 1950-1965. It was under permit from the Texas Water Quality Board for Petrochemical Waste Disposal as an industrial waste disposal facility until 1972. From 1966-1971, 70 million gallons of petrochemical wastes were disposed in the depressions that were created from sand extraction (an unlined 12-acre lagoon). In 1973, the permit was revoked. The primary contaminants in this waste were benzo(a)pyrene, vinyl chloride, and benzene. Benzene was at concentrations ranging from 400-500 mg/kg. Other primary contaminants in the sludge, soil, ground water, surface water, and air were volatile organic compounds (VOCs), phenols, heavy metals, and polychlorinated biphenyls (PCBs). These pose potential health risks including central nervous system disorders, liver damage, and cancer upon direct exposure.
In 1980, Superfund was enacted demanding immediate remediation of hazardous waste sites. In 1982, the EPA Superfund team, per request from the state department, constructed a large dike around the lagoon to protect from contamination spreading. This needed to be repaired later in the year. Some contaminated sludge was put back into the lagoon and some was placed in an approved landfill. In 1983, the French Limited was added to the EPA’s National Priorities List (NPL), which includes the most severe uncontrolled sites eligible for cleanup under the Superfund program.
In 1987, the EPA decided to try bioremediation, which was the first time that technology was used at a Superfund site.
Figure 12 shows the map of the French Limited Superfund Site
Figure 12 – French Limited Site Map (Environmental Protection Agency, 2013)
The site was extremely vulnerable because of its proximity to water sources and people. It is located within l00-year floodplain of the San Jacinto River and a shallow ground water system (20-50 feet deep) is in use by nearby residents. Figure 13 shows how flooding can spread the contaminants at the site. The nearest residence is within 300 feet of the main pit and the nearest drinking water well is within 1,500 feet of the main pit. This is a concern for the approximately 10,000 residents in Crosby and nearby communities. There are 300 residents within one mile of the site. Additionally, groundwater and surface water are used for drinking water and irrigation.
Figure 13 – Schematic of contamination at French Limited (Environmental Protection Agency, 1993)
Bioremediation was chosen because it offered a less expensive option to destroy the same amount of waste as an incinerator in the same amount of time.
In-situ slurry-phase bioremediation was conducted to remedy the site. Bioremediation of the lagoon sludge and water, surface water, and ground water began in 1992 after a year of technical design and construction. To aid the treatment process, pumps were used to mix lagoon sludge with lagoon liquids. First the bottom sludge is broken up and then contaminated soil beneath the sludge is brought up and mixed with the lagoon liquids. The process utilized microorganisms that were present at the site. After all the sludge and liquids in the lagoon was treated the lagoon water was sent to the water treatment facility on site. Figure 14 shows a schematic of the treatment system. Also after treatment, the soil was mixed with clean fill and the surface later used for vegetation. The ground water treatment was governed by VOC concentration. Monitoring will continue for 30 years after initial treatment. Treatment was done in two 17 million gallon cells. A technology known as Mixflo was used for aeration. This system minimized air emissions during the bioremediation process. The treatment process took 11 months and 300,000 tons of soil and sludge were treated overall.
To compensate damage to fouled habitat from lagoon spillage of hazardous waste, the French Limited Task Group (FTLG) planted 23 acres of new wetlands near the site.
Figure 14 – Schematic of treatment at French Limited (Environmental Protection Agency, 1993)
The concentration of benzene post-treatment was in the range of 7-43 mg/kg. Overall, the total cost, including treatment, pilot studies, technology development, and backfilling, was $49 million.
After initial remediation, the French Limited site has been revisited several times to mitigate contamination from floods. These floods caused spills and dispersed hazardous wastes into nearby areas. From 1980-1983, the EPA was involved in three remediation projects. In 1989, flooding affected the drinking water source. A wall was constructed to mitigate this issue and in 1994 when another flood occurred, the wall functioned successfully. In 1993, bioremediation was performed on the main waste lagoon. In 1996, a ten-year Natural Attenuation program began. The first Five-Year review was approved in 1995 and a second Five-Year review was approved in 2002. From 2002 to 2005, in-situ bioremediation was performed. The EPA conducted analyses and wrote Ground Water Monitoring Reports in 2007, 2008, 2009, 2010, 2011, and 2012.
Wheal Jane is an abandoned tin mine located in Cornwall, UK. As shown in Figure 15 it is located next to the Carnon River. In the winter of 1991/1992, just one year after closure, AMD from Wheal Jane contaminated the Carnon River (Whitehead & Neal, 2005a).
Figure 15 – Wheal Jane location (Whitehead & Prior, 2005b)
AMD is a big problem in the UK with South Wales, Yorkshire and Durham suffering the most. It is also an issue in popular mining areas of the US including Pennsylvania and eastern Appalachia (Whitehead & Prior, 2005b). A pilot study of bioremediation methods using passive wetlands was implemented and studied water pre-treated in three ways: Lime Dosing (LD), Anoxic Limestone Dosing (ALD) and Lime Free (LF). After treatment AMD was allowed to pass through aerobic cells and then into anaerobic cells. Figure 16 shows a schematic of the system (Whitehead & Prior, 2005b). This case study will look into the anaerobic cells only since it is the key technology that can be considered a bioremediation technique. The system at Wheal Jane is the largest of its kind in Europe and cost around £1 million ($1.6 million) to set up (Whitehead & Prior, 2005b).
Figure 16 – Schematic of treatment system at Wheal Jane (Whitehead & Prior, 2005b)
The anaerobic cells were 87.5x8.75x1 meters (length x width x depth) in dimension (Johnson & Hallberg, 2005b). The compost consisted of 95% sawdust, 5% hay and a small amount of manure to introduce bacteria to the cell. The compost was contained in polyethylene membranes. The main function of the cells was to generate alkalinity (i.e. to raise the pH) and to remove toxic heavy metals that included zinc, copper and cadmium by making use of SRB. Water that entered the anaerobic cell had already been pretreated in aerobic cells to remove iron and arsenic (Johnson & Hallberg, 2005b). Hydrolysis of the iron in the aerobic cells meant that influent into the anaerobic cells had a low pH approximately around pH 3. The cascading of water through the 5 aerobic cells also caused the water to have a high oxygen content. Both low pH and high oxygen content are known to be bad for SRB. Because of the large surface area lots of rainwater also entered the anaerobic cells.
Water was sampled in the rock filters and results of the study showed that (Johnson & Hallberg, 2005b):
Placing the anaerobic cells after the aerobic cells was found to be counterproductive, the water entering the anaerobic cells was found to have lower pH and higher oxygen concentrations that it would if it drained from the mine directly (Johnson & Hallberg, 2005b). Large surface area for the anaerobic cells was also found to be an issue since it allowed large volumes of rain water to enter the system. Around 0.2 l/s of water flowed through the cell of which AMD made up 56% and rainwater 44% (Johnson & Hallberg, 2005b). This severely limited the amount of water that could be effectively processed. Studying the soluble iron content of the LD and ALD cells showed that for the soluble iron coming from the AMD was 29% and 54% respectively. The remaining iron was found to be flushed from the sumps that connected cells. Removal of iron before it entered the anaerobic cell was found to be essential, since if it was allowed to enter the cell it would flow through, enter the rock filters and undergo hyrdrolysis and produce significant acidity (Johnson & Hallberg, 2005b).
The key factor that determined why the LF system worked but the LD and ALD did not had nothing to do with the influent water since for all three systems it was more or less identical. The main contributing factor was the fact that the LF system had to be shutdown for 10 months due to technical isssues. This downtime allowed the anaerobic cells in the LF system to “mature” (Johnson & Hallberg, 2005b). The harsh environment produced by the AMD likely killed any native bacteria in the manure but the shutdown time allowed resistant SRB to grow into sizable populations in the anaerobic cells (Johnson & Hallberg, 2005b).
Anaerobic cells should always be placed ahead of aerobic cells to raise the pH and lower the oxygen content of influent into the cell (Johnson & Hallberg, 2005b). This also limits the rain water entering the cells. The cells should be allowed to “mature” for a period of time before AMD is treated to allow robust SRB populations to develop in significant numbers (Johnson & Hallberg, 2005b). The optimum maturation period is still a topic of ongoing research. It is clear that bacteria from an innoculating material such as manure will not survive, it may be better to introduce a slurry from another “mature” anaerobic cell to boost the SRB population (Johnson & Hallberg, 2005b).
The Brown Wood Preserving Site (BWPS) is located 2 miles west of Live Oak, Florida. The site covers a 51-acre area in which sinkholes are a common geologic hazard due to the karst terrain (USEPA Office of Solid Waste and Emergency Response & Technology Innovation Office, 1995). The karst consists of limestone and dolomite estimated to be 2,500 ft thick. This limestone forms the upper region of the Floridian Aquifer which has an extensive network of connected cavities allowing quick movement of ground water. The site is not located in an area prone to flooding and the surrounding land is considered rural. Locally there are four water wells in private use, the closest located 1000ft south of BWPS. Public water supplies are drawn from wells that are 2 miles away and uphill.
BWPS operated for 30 years from 1948-1978 (USEPA Region IV, 1995) and (Sharma & Reddy, 2004). Pressure treatment of timber with creosote and occasionally pentachlorophenol was carried out at BWPS. Waste water produced in the process was treated and discharged to a lagoon. In 1981 a former site opertator informed EPA that the site had handled hazardous material and samples collected in 1982 found soil and sludge around the treatment tanks and lagoon to be contaminated with PAHs (USEPA Region IV, 1995). In 1982, the site was placed on the National Priorities List and in 1988 a Record of Decision (ROD) was signed (USEPA Office of Solid Waste and Emergency Response & Technology Innovation Office, 1995). The ROD identified a cocktail of 6 PAHs to be used as Total Carcinogenic Indicator Chemicals (TCICs) with a goal of 100mg/kg of TCICs within two years (USEPA Office of Solid Waste and Emergency Response & Technology Innovation Office, 1995). The PAHs selected were the 6 out of the 200 compounds that make up creosote that the EPA determined to pose the highest risk, these included benzo(a)anthracene, benzo(a)pyrene, benzo(b)fluorothene, chrysene, dibenzo(a,h)anthracene and indeno(1,2,3-cd)pyrene (USEPA Office of Solid Waste and Emergency Response & Technology Innovation Office, 1995). The EPA considered 7 different treatment options to be implemented at BWPS that included:
Land treatment was selected as the best remediation option based on cost and technical feasibility (USEPA Office of Solid Waste and Emergency Response & Technology Innovation Office, 1995). Although some of the sludge found initially at the site was pumped and transported to a hazardous waste landfill at Emelle, Alabama (USEPA Region IV, 1995).
The total concentration of the 6 TCICs in the soil varied from 100mg/kg to 208mg/kg. The soil conditions in the lagoon varied from clay to silty clay and fine sands (USEPA Office of Solid Waste and Emergency Response & Technology Innovation Office, 1995).
The EPA ROD required a number of goals be achieved. The total TCIC concentration had to be reduced to 100mg/kg within two years. The final concentration was selected so that the risk of soil ingestion by a child would be 1x10-6. Once the site was decontaminated the land treatment area had to be re-vegetated.
The land treatment system consisted of several components. A clay liner of thickness 1-3ft. A berm of compacted clay of height ranging from 2.5ft-7ft surrounding the lagoon and an additional berm surrounding the stockpiled soil of height 3ft. Water drains to prevent surface water entering the site. A separate subsurface drainage system under the treatment area to collect runoff water. A 750,000 gallon retention pond to contain any runoff water. A sprinkler system able to deliver water at 0.5inches an hour. Figure 17 shows the site layout.
Figure 17 – Site layout of the land treatment system (USEPA Office of Solid Waste and Emergency Response & Technology Innovation Office, 1995)
The land treatment area (LTA) had an area of four acres and was split into 8 half-acre treatment plots. The treatment was completed in 3 lifts with a sample from each plot being collected periodically; when the TCIC concentration for that lift fell bellow 100mg/kg a new lift was treated (USEPA Office of Solid Waste and Emergency Response & Technology Innovation Office, 1995).
The first lift was placed in January 1989 and consisted of 3300 yd3 of soil placed in a layer 5-7 inches thick. The LTA was treated twice a week with microorganisms capable of biodegrading the PAH TCICs in the contaminated soil. The LTA was tilled once every two weeks and the water content of the soil was maintained at 10% using the sprinkler system (USEPA Office of Solid Waste and Emergency Response & Technology Innovation Office, 1995). Initial TCIC concentrations varied from 103-258mg/kg. After 8 months of treatment the concentration of TCICs in each plot was reduced to 100mg/kg or less. Removal rates varied from 58mg/kg a month to 13mg/kg a month.
The second lift was placed in September 1989. Around 3000 yd3 was placed in a layer 9-12inches thick. Again the same treatment and tilling process was followed. Initial TCIC concentrations were similar to those in lift 1 and within 3 months the TCIC concentrations were below the 100mg/kg target.
The final lift was 1800 yd3 placed in a layer 4-7inches thick. The same treatment and tilling process was followed. Final TCIC concentrations after treatment was completed in July 1990 varied from 23mg/kg to 92mg/kg. Once remediation was completed 90% of the LTA was successfully covered with native grass species. The final cleanup goal was achieved in 18 months, 6 months ahead of the ROD 24 month deadline (USEPA Office of Solid Waste and Emergency Response & Technology Innovation Office, 1995).
Costs for similar projects conducted in the past varied from $50-$100 per cubic yard of soil treated, provided a minimum volume of 3000 yd3 was treated. This cost of soil treatment was a total of $565,406 to treat 8100 yd3 of soil, meaning the cost per volume was $70 per cubic yard. Cost breakdown showed that of the total cost $58,039 was for ‘before treatment’ operations and $9,827 was for ‘post treatment’ operations.
(USEPA Office of Solid Waste and Emergency Response & Technology Innovation Office, 1995).