Editors: Gra 3. Principles of bioremediation...

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Trends in Bioremediation and Phytoremediation, 2010: 23-54 ISBN: 978-81-308-0424-8 Editors: Grażyna Płaza

3. Principles of bioremediation processes

Artin Hatzikioseyian National Technical University of Athens (NTUA), School of Mining and Metallurgical Engineering, Division of

Metallurgy and Materials Engineering, Laboratory of Environmental Science and Engineering Heroon Polytechniou 9, 15780 Zografou, Athens, Greece

Abstract. The field of bioremediation has experienced a dynamic evolution and remarkable development over the past decades. The bioremediation of contaminated former industrial sites or waste deposits became a well established field of environmental biotechnology. This chapter gives an overview of engineered bioremediation treatment options and focuses mainly on the biological aspects of biostimulation, bioaugmentation, monitored natural attenuation (MNA), and biotransformations of metals, metalloids and radionuclides in contaminated sites. Some aspects of phytoremediation are also presented.

Introduction Bioremediation is concerned with the biological restoration and rehabilitation of historically contaminated sites and with the cleanup of areas contaminated in more recent times, either accidentally or incidentally, as a result of the production, storage, transport, and use of organic and inorganic chemicals [1, 2]. Bioremediation offers the possibility of degrading, removing, altering, immobilising, or otherwise detoxifying various chemicals from the environment through the action of bacteria [3-5], fungi [3, 4] and plants [5-11]. Most of the advances in bioremediation have been realised through the assistance of the scientific areas of microbiology, biochemistry, molecular biology, analytical chemistry, chemical and environmental engineering, among others. These different fields, each with its own individual approach, have actively contributed to the development of bioremediation progress in recent years [12]. Correspondence/Reprint request: Dr. Artin Hatzikioseyian, National Technical University of Athens (NTUA), School of Mining and Metallurgical Engineering, Division of Metallurgy and Materials Engineering, Laboratory of Environmental Science and Engineering, Heroon Polytechniou 9, 15780 Zografou, Athens, Greece. E-mail: [email protected]

Artin Hatzikioseyian 24

The systematic use of biological processes in the design of engineered environmental remediation systems is today well developed and documented [1, 4, 12-44]. Contemporary environmental biotechnology utilises techniques that employ usually natural strains of microorganisms in order to mobilise and remove organic or immobilize inorganic pollutants from the environment. The key advantage of bioremediation processes as compared to other biological technologies is that they can employ enzymatic metabolic pathways that have been evolved in nature over very long periods of time, thus becoming very specific. The combination of such pathways can make possible the degradation of a wide variety of even hazardous pollutants. Today, metabolic pathways for the degradation of compounds previously considered as non degradable have been identified [15]. Laboratory and field techniques for assessing the applicability of biological processes for the degradation of the pollutants that are present at specific sites have become available [25, 45, 46]. Historically the engineering aspects of environmental biotechnology were perceived as an extension of the conventional biological wastewater treatment processes. The idea was to bring together microorganisms and the waste organic materials under conditions that would enable the microorganisms to utilise the organic molecules as a substrate. The corresponding process efficiency could then be measured by the use of an easy lumped parameter such as BOD5 or COD and occasionally by the use of compound specific analyses. Experience however quickly led to the realisation that systems dealing with environmental contamination through bioremediation were much more complex, and that moving from the flask to the field was not a simple task. Today, the experience accumulated over the last decades, has improved our understanding in many aspects of this multidisciplinary technology. However, technical and non technical issues associated with large scale implementation of bioremediation technology still exist. The non technical level of uncertainty associated with the technology is the result of combination of legal, regulatory, social and financial considerations. Non technical issues can at times override the potential technical feasibility for a specific application. The combined evaluation of the technical and non technical issues is an important step towards the successful application of environmental biotechnology in remediation. The state of the art of bioremediation technology as well as examples of more or less successful case studies are published in many books during the last 20 years [14, 16, 19, 25, 26, 29, 31, 35, 37, 39-41, 43-45, 47]. Many books are specialized in specific applications of bioremediation such as bioremediation of MTBE [48], perchlorate [49], oil spills – hydrocarbons contamination [29, 39, 50], chlorinated solvents [51] metal ions [27, 29, 34, 52-54], radionuclides [38, 55], explosives [56] in various environments, e.g. marine harbours [57], cold areas [58-60] etc. The financial aspects of bioremediation is also very important and documented [23, 44]. Finally, a novel area such as phytoremediation has gained attention in the recent years [5-10, 29]. Currently many websites are also providing researchers with advice tools, technologies and methods for assessing and cleaning up hazardous waste sites. The hazardous waste clean-up information (CLU-IN) website (http://clu-in.org/) provides information about innovative treatment and site characterization technologies to the hazardous waste remediation community. It describes programs, organizations, pub-lications, and other tools for federal and state personnel, consulting engineers, technology developers and vendors, remediation contractors, researchers, community groups, and individual citizens. The site was developed by the US Environmental Protection Agency (EPA) but is intended as a forum for all waste remediation stakeholders. The Federal

Principles of bioremediation processes 25

Remediation Technology Roundtable (FRTR) provides the remediation technologies screening matrix and reference guide version 4.0, which serves as "yellow pages" of remediation technologies. It is intended to be used to screen and evaluate candidate cleanup technologies for contaminated installations and waste sites in order to assist remedial project managers (RPMs) in selecting a remedial alternative. To reduce data collection efforts and to focus the remedial evaluation steps, information on widely used and presumptive remedies is provided [61]. The guide is available online from the website (http://www.frtr.gov/matrix2/top_page.html). Some more web resources are presented and discussed in the literature [62, 63]. It is beyond the scope of this chapter to cover every aspect of bioremediation in details and to review any relevant results which have been obtained by the huge community of researchers in science and industry in the last decades. However, a general overview is attempted and a collection of interesting books and selected papers from the literature is proposed for in depth study. Contaminated soil remediation options In terms of managing soil pollution, there are three generic approaches. The first approach is to contain the contamination within the affected area. This option does not reduce the total concentration of the pollutants in soil, but aims to manage exposure to the pollutants, and hence environmental and health risks, through technologies that lower pollutant mobility and bioavailability. The second approach is to treat the contaminated mass in-situ, while the third approach is to remove the contaminated mass from its surrounding environment and then treat it or dispose it ex-situ. The last two clean-up technologies aim to reduce significantly the total concentration of the pollutants in soil to a maximum allowable total concentration by applying passive or active bio/remediation technologies [64]. Containment systems Containment is necessary whenever contaminated materials are to be buried or left in place at a site. The containment option for soil, sediment, bedrock and sludge includes technologies such as cover systems, vertical barriers, horizontal barriers and hydraulic control measures [55]. Cover systems: A covered system is composed by a multilayer construction placed over the contaminated soil to reduce the harmful effects of the contaminants at the surface, minimise water infiltration through the contaminants by rain, prevent upward migration of groundwater by capillary rise, prevent airborne migration of the contaminants, and where appropriate control gas migration (Figure 1). With such systems contaminated soil and ground water can be physically isolated with low-permeability barriers such as landfill caps, liners, and cutoff walls. The optimum combination of the layers, in terms of composition, thickness, and sequence of materials is based on an assessment of the physical and chemical properties of the entire system (e.g., chemical resistance, physical resistance to climatic conditions, and ground conditions such as cracking and channelling due to drought, freeze/thaw, settlement), construction aspects, consideration of the reduction in environmental risk of the underlying contaminated land and cost [55]. Construction difficulties and questions about long-term reliability have limited the application of containment technologies in the past. Questions such as: (a) How can methods for detecting defects in containment systems be improved? (b) How


Figure 1. Schematic layout of covered system [61]. can the bottoms of vertical walls be effectively sealed? (c) What is the long-term reliability of different materials used for containment? and (d) How significant is diffusive transport of contaminants across barriers over long time scales? always arise [65]. Technical details concerning the characteristics of geomembranes and the construction of covered systems can be found in the literature [61, 64, 66, 67]. Vertical barriers: Vertical barriers are installed adjacent to contaminated ground (i) to prevent the off-site lateral migration of contaminated groundwater, (ii) to divert clean groundwater away from contaminated ground, and (iii) to reduce the extraction rates of contaminated groundwater from hydraulic control measures. They can be used to funnel groundwater to an in-ground treatment center (so called funnel and gate) and also be used to cut-off the underground migration of gases. To be effective, vertical barriers are normally tied into a natural low permeable layer at depth (e.g., a clay layer) or to an in-ground horizontal barrier. There are three common types of vertical barriers (i) displacement systems (e.g., sheet piling, membrane walls), (ii) excavated barriers (e.g., shallow cut-off walls, slurry trench walls, secant walls), and (iii) injection barriers (e.g., chemical grouting, auger mixing, jet grouting). Horizontal barriers: In-ground horizontal barriers are installed below the contaminated ground to prevent vertical migration. They can be used in combination with vertical barriers to isolate potentially mobile contamination. Horizontal barriers are generally formed by injection of cement slurries at depth [55]. Hydraulic control measures: Hydraulic control measures are used to adjust the groundwater flow around a contaminated area so that no further spread of contamination takes place. This can involve preventing or reducing the contact of the groundwater with the contaminated mass (e.g., lowering the water table), reducing, intercepting or containing a plume of contaminated groundwater, supporting other remediation methods such as in-ground barriers, or being part of groundwater remediation operations. Hydraulic control measures are commonly carried out by pumping out groundwater from a number of wells, or using diversion or interceptor trenches [55, 68]. Immobilisation, stabilisation and solidification of the contaminated area either in-situ or ex-situ aim to reduce the mobility of the contaminants by the following actions:

Stabilisation: Forming chemically immobile compounds of the contaminant, Solidification: Binding the soil together to form a monolithic block to prevent access by external mobilising agents such as wind, rain, and groundwater. Contaminated


soil can be contained by solidifying it in place with chemical fixatives (e.g. cementing agents including pozzolan-portland cement, lime-flyash pozzolan, and asphalt) [65], Vitrification: Melting and rapidly cooling the soil so that the contaminants are immobilised and encapsulated into a glassy matrix. Vitrification may be carried out in-situ or ex-situ. In-situ vitrification is performed by introducing an array of four graphite electrodes into the soil and heating them electrically with powerful generators to temperatures between 1600–2000 °C. At these temperatures the soil melts and forms a glass block. Upon cooling, organic contaminants are pyrolysed and reduced to gases during the melting process, while heavy metals remain enclosed in the stabilised glass mass. This method has also been successfully used in treating soils contaminated by radioactive materials. Vitrification may also be done ex-situ in special appliances where contaminated soil would be molten in presence of borosilicate and soda lime to form a solid glass block [43, 55].

In-situ treatment systems In the case of in-situ techniques, soil and associated groundwater are treated in place without excavation, [69]. Examples of in situ techniques include pump and treat, percolation (flooding), bioventing, air sparging, bioslurping and permeable reactive barriers, [23-25]. In-situ remediation ranges from partially closed (contained) systems to completely open ones such as oil spills, (i.e. Exxon Valdez case [70]). Pump-and-treat systems, which are applied to saturated-zone remediation the removal of any contaminated water from the ground, treatment either at an on-site or off-site plant and return it back to the contaminated soil zone, (Figure 2) [64, 69, 71, 72]. This is one of the most traditional methods of remediation, but it is not the most effective method for all contaminants because it is costly in investment - maintenance and very time consuming [72]. However, the advantage of this method is that the contaminant is actually removed entirely from this system if it is water soluble. For example in the cases of methyl tertiary-butyl ether (MTBE), which is a fuel additive very soluble in water, chlorinated solvents and hydrocarbons in groundwater, the pump and treat method combined with carbon adsorption and air stripping is very effective [72, 73]. However, for contaminants that bind very closely to the soil, such as polycyclic aromatic hydrocarbon (PAHs), desorption of the contaminant from the soil to the groundwater is very slow. When the groundwater is pumped out of the system, the contaminant present in the aqueous phase is also removed, but a significant portion of the contaminant still remains present in the ground. In order for pump and treat to be effective, it must be done over a long period, in order to give the contaminant sufficient time to desorb from the soil. However, this option becomes much more cost efficient and timely with the addition of mobility agents such as cosolvents (e.g. alcohols), surfactants (e.g. amphiphilic polyurethane nanoparticles, extracellular polymers, and cyclodetrins) and steam, that will loosen the bond of the contaminant to the soil particles and increase the apparent solubility of the contaminant [69, 74-76]. Some of the more conventional modifications of groundwater pump-and-treat remediation and aquifer restoration strategies from petroleum hydrocarbons include bioenhanced degradation [72]. Bioenhanced biodegradation can be done in situ and involves the introduction of nutrients and air or peroxide to enhance the natural biologic processes. Aboveground treatment can include a variety of strategies including chemical precipitation, evaporation, reverse osmosis, and ion exchange to remove metals, and carbon


Figure 2. A simplified pump and treat scheme [77]. Treated water is usually returned back in the ground water. absorption, air stripping, ultraviolet oxidation and use of bioreactors to remove organics. Common dissolved hydrocarbon constituents treated by these methods are the aromatic fuel hydrocarbons such as benzene, toluene, ethylbenzene, xylenes and fuel additives such as lead and methyl tertiary-butyl ether (MTBE) [72]. Percolation consists of applying water, containing nutrients and possibly a microbial inoculum, to the surface of a contaminated area and allowing it to filter into the soil and mix with the ground water [69]. Bioventing is the process of injecting air (i.e., aeration) to an unsaturated soil zone through the installation of a well(s) connected to associated pumps and blowers which draw a vacuum on the soil [17, 58, 63, 69]. This process combines an increased oxygen supply with vapour extraction. A vacuum is applied at some depth in the contaminated soil which draws air down into the soil from holes drilled around the site and sweeps out any volatile organic compounds. The development and application of venting and bioventing for in situ removal of petroleum from soil have been shown to remediate approximately 800 kg of hydrocarbons by venting, and approximately 572 kg by biodegradation [18, 78]. Air sparging or biosparging involves the injection of air into the saturated zone of a contaminated soil, at low flow rates (<5 m3/h per point) [17, 58, 69]. This is used to increase the biological activity in soil and to promote aerobic biodegradation by increasing the O2 supply via sparging of air or oxygen into the soil (Figure 3). In some instances air injections are replaced by pure oxygen to increase the degradation rates. However, in view of the high cost of this treatment in addition to the limitations in the amount of dissolved oxygen available for microorganisms, hydrogen peroxide (H2O2) is introduced as an alternative, and it is used on a number of sites to supply more oxygen [18]. In biosparging volatilization is typically less than that of the standard air sparging system [58]. Bioslurping is an in situ technology that combines vacuum-enhanced free-product recovery with bioventing of subsurface soils to simultaneously remediate petroleum


Figure 3. Biosparging system used with soil vapor extraction [63].

Figure 4. Bioslurping technology [63]. hydrocarbon-contaminated groundwater and soils (Figure 4). Vacuum-enhanced recovery utilizes negative pressure to create a partial vacuum that extracts free product and water from the subsurface. The technology is portable and uses a single pump to extract free product, groundwater, and soil gas from multiple wells. Groundwater and soil gas may require treatment before being discharged. Bioslurping is used at petroleum spill sites and has proven most effective in fine-to-medium textured soils or fractured rock in areas with a low water table [44, 72].


Ex-situ treatment systems In the case of ex-situ techniques, soil and groundwater are removed from their original locations for treatment. Examples of ex-situ techniques include land farming, irrigation, compost piles, engineered biopiles, and ex-situ slurry techniques [69]. Land farming is land treatment of soil for degradation or transformation of contaminants by a combination of volatilization and biodegradation by indigenous microorganisms. A common practice is to place the soil as a shallow layer within a bermed and lined treatment cell (or biocell), occasionally amend the soil with nutrients and water to stimulate biodegradation, and regularly till (aerate) the soil to mix and aid contaminant volatilization [17, 58, 69]. A land farming bioremediation case study for a site contaminated with hydrocarbons (e.g. PAHs, BTEX ) is presented in the literature [15]. Compost piles consist of soil supplemented with composting material (i.e. wood chips, straw, manure, rice hulls, etc.) to improve its physical handling properties and its water- and air-holding capacities. Although compost piles are exposed to the atmosphere, the interior is often anaerobic due to the oxygen demand of the contaminants and amendments. Thus, air is drawn through the compost (by vacuum, although aerated piles have been used to enhance the drainage) to supply O2 to the soil for promoting aerobic degradation of organic material and remove evaporated water. Compost piles are subjected to intermittent mixing using specially manufactured equipment that is capable of turning the pile over onto itself. Temperatures can increase to 60–70°C due to the exothermic nature of biodegradation, and mixing, aeration, and moisture addition help dissipate excess heat that could be inhibitory to biodegradation [15, 69]. One advantage of composting is that it is more effective than other solid-phase treatment systems for soils and sludges contaminated with viscous substances such as coal tar, creosote, or petroleum production and still bottoms. Soil treatment using composting systems is limited to biodegradable chemicals. The technology cannot treat metals and most other inorganic chemicals. Additionally, the technology cannot readily biodegrade halogenated chemicals. However the composting system effectively remediates soils that are heavily contaminated and cannot be treated by in-situ methods as well as wastes containing hazardous volatile constituents untreatable by land farming methods [44]. An example case study of composting for bioremediation of a site contaminated with explosives (e.g TNT, RDX) is presented in the literature [15]. Biopile is a remediation technique that involves placing contaminated soils into piles or cells above ground, and stimulating aerobic or anaerobic microbial activity within soils through controlled aeration and/or addition of minerals and nutrients (Figure 5.). Air is supplied to the biopile via a pipe-and-pump system, which either forces air into the pile (positive pressure) or draws air through the pile (negative pressure). Forcing air into the pile helps maintain constant temperature and aerobic conditions, while drawing air out of the pile can create anaerobic conditions. Although composting systems require large amounts of nutrients and bulking agents, fewer additives are needed for biopiles. Biopiles are normally operated at lower temperatures since less organic material is added. Biopiles have some potential limitations. For example, certain chemicals such as polychlorinated biphenyls (PCBs) and other hydrocarbons are resistant to biodegradation. In addition, high concentrations of toxic metals, such as lead, copper, and mercury, may limit biopiles treatment [44, 58, 63, 69]. Ex situ slurry techniques involve the creation and maintenance of a soil-water slurry as the bioremediation medium. The slurry can be maintained either in a bioreactor or in a pond or lined lagoon [29, 69]. Adequate mixing and aeration are key design requirements


Figure 5. Typical biopile system [63]. for slurry systems. Decomposition of organic contaminants takes place usually via aerobic processes. Nutrients and, perhaps, inoculum may be added to the slurry [69]. A significant benefit in the use of slurry biodegradation is the enhanced rate of contaminant degradation, a direct result of improved contact between the microorganisms and hazardous compounds. The agitation of contaminants in the water phase provides high degree of solubilisation of compounds and greater homogeneity. More details for slurry biodegradation technology can be found in the literature [16, 29]. Ex-situ solid applications involve the addition of water, nutrients and sometimes addition of cultured indigenous microbes or inocula. They are often conducted on lined pads to ensure that there is no contamination of the underlying soil [69]. Both in-situ and ex-situ techniques are capable of saturated and unsaturated zone remediation, although restriction exists depending on the exact system used. However, many factors can influence the effectiveness of each technique [69]. Ex situ techniques allow more opportunities to control or engineer conditions for remediation. Although some sites may be more easily controlled with ex-situ configurations, others are more effective with in-situ treatment. For example, many sites are located in industrial and commercial areas, making excavation extremely difficult. In addition if the contamination is deep in the subsurface, excavation becomes too expensive. As a result of these physical barriers, the required excavation efforts may make ex-situ biotreatment impracticable. Other factors could also have an impact on the type of treatment. At a typical site, the contamination is basically trapped below the surface. Exposing the contamination to the open environment through excavation can result in potential health and safety risks. In addition, the public’s perception of the excavation of contaminants could be negative, depending on the situation. All of these conditions clearly favour in-situ biotreatment. Nonetheless, the key is to carefully consider the parameters involved with each site before evaluating which technique to use [74]. Factors affecting bioremediation The most important principle of bioremediation is that microorganisms (mainly bacteria or fungi) can be used to destroy hazardous contaminants or transform them to less harmful forms. Thus, bioremediation of contaminants is an application of the


microbial metabolic activity. Microorganisms, through their enzymatic pathways, act as biocatalysts and facilitate the progress of biochemical reactions that degrade the targeted contaminants. As a result, bioremediation techniques are only applicable in environments that can sustain life. The microorganisms act against the contaminants only when they have access to a variety of materials-compounds to help them generate energy and nutrients to build more cells. In very few cases the natural conditions at the contaminated site provide all the essential materials in large enough quantities that bioremediation can occur without human intervention - a process called intrinsic bioremediation. More often, bioremediation requires the construction of engineered systems to supply microbe-stimulating materials - a process called engineered bioremediation. Engineered bioremediation relies on accelerating the desired biodegradation reactions by encouraging the growth of more organisms, as well as by optimizing the environment in which the organisms must carry out the detoxification reactions [47]. The metabolic characteristics of the microorganisms in association with the physicochemical properties of the targeted contaminants determine whether a specific microorganism - contaminant interaction is possible. The actual successful interaction between the two, however, depends on the environmental conditions of the site of the interaction. Specific constrains should therefore be fulfilled for a successful bioremediation attempt. These constrains encompass the microbial, chemical and environmental characteristics of the targeted site. Microbial constrains A successful bioremediation effort relies on the utilisation of the appropriate microorganisms [25, 79]. Such microbial populations can in theory be consortia of naturally existing species or genetically engineered microorganisms. Most applications rely on the use of naturally existing microbial populations which often are not well characterised. That is to say the microbial populations are effective in their desired application but the complete characterisation of the population is not well known. This knowledge gap is not necessarily the result of a scientific inability but rather of the continuous dynamic adaptation of the microbial species to their environments. An example of this ability of microbial populations to adapt to the presence of man-made chemicals comes from the field of medicine, where the rapid adaptation of pathogenic organisms and their resulting immunity to specific classes of antibiotics as a result of the excessive use of these antibiotics has been well documented. These adaptational mechanisms advance through selection processes in which variant species with a specific survival advantage for the given environment take over and survive successfully. The survival advantage often relies on the ability of an organism to metabolise as substrate organic molecules (pollutants) existing in a given site. Contemporary microbiological techniques allow the identification of such transconjugants that originate from a background microbial population confirming that such processes are active in bioremediation practice [70]. Horizontal transfer of catabolic plasmids among different species existing within a site may also result into species that possess enhanced catabolic or resistance potential [80]. Such plasmid containing bacteria have been isolated from polluted sites [81-84]. Species that can through such plasmid transfer catabolise as sole carbon source hazardous xenobiotic compounds (as for example 3-chlorobenzoate) have been reported [85, 86]. The ultimate impact, however, of such plasmid transfer processes on the field application potential of bioremediation will have to pass through the previously described path of


natural selection. A newly acquired metabolic advantage will be assessed, through the mechanism of natural selection, and may allow the ultimate successful establishment of a transconjugant species in a contaminated site. Genetically modified microorganisms (GMOs) have often been presented as offering a major potential advantage for bioremediation. The development of recombinant DNA and other genetic engineering technologies, in the late 1970s, was believed that could be widely applied for environmentally-beneficial purposes, including the clean-up of contaminated soil and water [80, 87-89]. The continuously growing knowledge on catabolic pathways and critical enzymes provides the basis for the rational genetic design of new and improved enzymes and pathways for the development of more effective processes. Many researchers had expected that genetically modified organisms having novel biochemical traits or enzymatic activities would quickly find broad applicability in bioremediation of hazardous chemicals from the environment [20, 30]. However the practical impact of GMOs is likely to remain low for many key reasons. Public, regulatory, economic and technical issues associated with the release of genetically engineered, or recombinant, microbial species into an open environment usually arise. Many site owners, consultants and regulators are more comfortable choosing technologies and methods with which they are familiar, have a long track record of success and thus a greater predictability. Legislative reasons are associated with the strict control on the release of such organisms into the environment. There is significant concern about the long term survival of genetically engineered species into a natural environment where they would have to compete with the naturally existing consortia that had ample time to adapt to the prevailing environmental conditions. Thus, difficulties in obtaining permission to use genetically engineered microorganisms from government regulatory agencies as well as public controversies have made companies reluctant to develop bioremediation strategies based on GMOs [20, 30]. Finally, their use is considered costly. Technically speaking, it seems more plausible to use GMOs in ex-situ bioremediation treatment schemes in bioreactors, designed for use with defined soil slurries or water streams in tightly controlled environments. Not only does this limit the widespread release of the GMOs into the environment and avoids the problem of competition with indigenous microflora, but also allows the microorganism to be maintained at controlled temperatures and other growth conditions, to be used with relatively well-defined waste streams containing one or a small number of specific contaminants. The application of the genetically engineered microorganisms in industrial scale bioremediation is not yet prominent. Until today genetically engineered microorganisms have not been used in commercial site remediation projects, with few only exceptions [90]. Most bioaugmentation projects have used naturally-occurring bacteria for which obtaining regulatory approval is relatively easy. However, recently transgenic plants begin to find applicability in commercial phytoremediation projects. Chemical constrains

Bioavailability of contaminants In order for the pollutants to be amenable to biological degradation they must be bioavailable [91]. Bioavailability is related to the physical state of the contaminant and the possibility of efficient contact between the microorganism and the contaminant. This contact is best when the microorganism-contaminant interface is maximised. Regarding physical state, microorganisms generally assimilate pollutants from the liquid phase and


cannot effectively degrade a pollutant until it desorbs from aquifer solids, diffuses out of nanopores, or dissolves from nonaqueous phase liquids (NAPLs) into the bulk solution. In such cases, the rate of biodegradation can be controlled by the diffusion, desorption, or dissolution rates. Polar, water soluble contaminants are more easily bioavailable. The increase of the contaminant - microorganisms contact surface for hydrophobic contaminants may require the addition of surface active agents. Knowledge of partitioning and rates of transfer of a chemical between its disolved-sorbed-volatile states becomes important in defining its bioavailability. Bioavailability comprises the effects of all the physical and chemical parameters that eventually dictate the potential for the microbial utilisation of a compound and thus its biodegradation potential [15, 33, 46, 91]. Biodegradability of contaminants The success of any bioremediation project depends mainly on the chemical structure of the organic molecules present in the contaminated site [25, 46]. Some structural features of organic compounds that are not common in nature, called ‘‘xenophores’’ (e.g., substitutions of H with Cl, NO2, CN, and SO3 groups), make such molecules difficult to be metabolized by microorganisms. Thus, contaminants that contain such xenophores tend to be recalcitrant to microbial degradation, [14, 25, 33, 46]. Table 1. presents the experienced biodegradability potential of different target organic molecules. Numerous mechanisms and pathways have been elucidated for the biodegradation of a wide variety of organic compounds [25]. All metabolic reactions are mediated by enzymes. These belong to the groups of oxidoreductases, hydrolases, lyases, transferases, isomerases and ligases. Many of the oxygenase enzymes that attack aromatic hydrocarbons have a remarkably wide degradation capacity due to their non specific substrate affinity. For example, toluene dioxygenase is capable of degrading more than 100 different compounds, including TCE, nitrobenzene, and chlorobenzene. Other examples are esterases, which break down ester bonds by the addition of water; depolymerases, which hydrolyze polymers; dehalogenases, which remove halogen atoms

Table 1. Biodegradability of various compounds [37, 63].

Simple hydrocarbons, C1–C15 Very easy

Alcohols, phenols, amines Very easy

Acids, esters, amides Very easy

Hydrocarbons, C12–C20 Moderately easy

Ethers, monochlorinated hydrocarbons Moderately easy

Halogenated and non-halogenated volatile organic compounds (VOCs) Moderately easy

Halogenated and non-halogenated semi-volatile organic compounds (SVOCs) Moderately easy

Hydrocarbons, greater than C20 Moderately difficult

Multichlorinated hydrocarbons Moderately difficult

Polycyclic aromatic hydrocarbons (PAHs), Polychlorinated biphenyls (PCBs),

Pesticides and herbicides

Very difficult


such as chlorine and replace them with —OH groups; decarboxylases which remove CO2 groups (i.e., decarboxylation), hydratases which add water to alkenes converting them into secondary alcohols; glutathione S-transferase which transfers the thiol group to chlorinated compounds with concomitant dechlorination; racemases which catalyze L- and D-amino acid interconversions and finally CoA-ligase, which adds -S-CoA to fatty acids during beta-oxidation. For details of different biotransformation pathways the reader is referred to the literature [15, 22, 25, 46, 92]. Other contaminant properties Contaminant properties are critical to contaminant-soil interactions, contaminant mobility and to the ability of treatment technologies to remove, destroy or immobilize contaminants. Important contaminant properties include: Solubility in water, dielectric constant, diffusion coefficient, molecular weight, vapour pressure, density and aqueous solution chemistry [40]. Nutrients Most in-situ bioremediation methods practiced today rely on the stimulation of indigenous microbial populations at the site of contamination, by addition of appropriate nutrients, principally carbon, oxygen, nitrogen and phosphorus, and by maintaining optimum conditions of pH, moisture and other factors, to trigger increased growth and activity of indigenous biodegradative microorganisms [17]. Nitrogen and phosphorus requirements are often estimated by calculating a carbon to nitrogen to phosphorus ratio C/N/P close to 100/(10 to 5)/1. Many authors report optimum experimental results C/N/P ~70/3/0.6, [93], 8/1/0.07, [94], 1/11/2 [95] for crude oil bioremediation of different origin. Fertilizers such as paraffinized urea and octylophosphate in C/N/P 100/10/1 respectively have been suggested for optimal growth. Dibble and Bartha [96] suggest ratio of C/N/P 800/13/1, illustrating that the nutrient requirement is specific to oil-in-water mixtures and needs individual consideration for any case. Suggested C/N values for composting are between 30-40 [97]. A detailed excellent review for nitrogen and phosphorous requirements for bioremediation as well as the detrimental effects of excess nutrients can be found in the literature [98]. Nutrients typically are delivered by controlling ground water flow using injection wells or burred perforated pipes (infiltration gallery). In the most common configuration, ground water withdrawn from production wells downgradient from the biostimulation zone is amended with the nutrients required for biostimulation, treated if necessary to remove contaminants, and reintroduced to the aquifer upgradient of the biostimulation zone using the injection wells or infiltration galleries. Water from an external source is required if the flow of withdrawn water is insufficient to control the subsurface flow or if it is infeasible to reinject the withdrawn ground water. The rate of nutrient delivery to the biostimulation zone, therefore, is often limited by the solubility of the nutrients in water and the reinjection flow rate [47]. Oxygen, air, hydrogen peroxide In the majority of applications, bioremediation is an oxidation process. During oxidation of the contaminants, microorganisms extract energy via electron transfer. Electrons are removed from the contaminant and transferred to a terminal electron acceptor which, during aerobic biodegradation, is oxygen. During decomposition of the


organic substrate, oxygen concentrations in the subsurface may become depleted [29]. The major kinetic limitation on aerobic bioremediation is often the availability of oxygen due to the low solubility of oxygen in water. This is more intense in the cases of organic molecules with high oxygen demand such as petroleum hydrocarbons. Air, oxygen, or other oxygen sources (e.g., hydrogen peroxide, ozone) may need to be added to the infiltration water to promote aerobic biodegradation. Air sparging of water can supply 8 mg/L dissolved oxygen, sparging with pure oxygen can deliver 40 mg/L, while application of hydrogen peroxide can provide more than 100 mg/L oxygen. Therefore, while air sparging is the simplest and most common oxygen delivery technique, the use of oxygen or hydrogen peroxide may speed the bioremediation process and decrease the pumping required. However, in some cases the increased cost and potential explosion hazard associated with pure oxygen supply may limit the applicability of direct oxygen use [47]. On the other hand, application of hydrogen peroxide to in-situ bioremediation is limited by its toxicity to microorganisms, its potential for causing aquifer plugging due to the highly reactive nature of hydrogen peroxide resulting in chemical oxidations of organic and inorganic compounds, producing precipitates [99], and the formation of oxygen bubbles which decreases aquifer permeability [16, 100, 101]. The problems associated by the use of alternative oxygen sources are discussed in the literature [29]. Alternative electron acceptors In the absence of molecular oxygen, anaerobic microorganisms use other forms of combined oxygen. For example, denitrifying bacteria use nitrate (NO3

-), nitrite (NO2-), or

nitrous oxide (N2O); dissimilatory metal-reducing bacteria use manganese or ferric iron oxides (e.g., MnO2, Fe(OH)3, or FeOO-); sulfate-reducing bacteria use sulfate (SO4

2- ); and methanogens use carbon dioxide (CO2) or bicarbonate (HCO3

-) as electron acceptors [15, 102]. In cases where oxygen is progressively depleted, electron acceptors are generally used up in a set sequence determined by the appropriate redox potentials of the oxidation reactions under consideration [15, 103]. Thermodynamic concepts imply the following sequence of electron acceptor utilization:

4+ 3+ 22 3 4 3O NO Mn Fe SO HCO− − −→ → → → →

The implication of this thermodynamic analysis is that when the electron acceptor demand is relatively high (e.g., near the source zone), microbial degradation would sequentially deplete the available oxygen, then nitrate, manganese, ferric iron, and sulfate before methanogenesis becomes predominant. Thermodynamic considerations also imply that heterotrophic microorganisms capable of deriving the maximum amount of energy per unit of carbon oxidized would have a competitive advantage over other species, and their respiration mode would become dominant until their specific electron acceptor is used up [15]. Metal ions Although some metals are essential in trace quantities for microbial growth, heavily contaminated sites with high concentrations of metal ions in contaminated soil or water usually inhibit the metabolic activity of the cells, thus affecting directly any bioremediation process [74].


Toxic compounds High aqueous phase concentrations of some contaminants can create toxic effects to microorganisms, even if the same chemicals are readily degraded at lower concentrations. Toxicity prevents or slows down microbial metabolic activity and often prevents the growth of new biomass needed to stimulate rapid contaminant removal. The degree and mechanisms of toxicity vary with specific toxicants, their concentration, and the exposed microorganisms. Some organic compounds are toxic to targeted life forms such as insects and plants and may also be toxic to microbes. These compounds include herbicides, pesticides, rodenticides, fungicides, and insecticides. In addition, some classes of inorganic compounds such as cyanides and azides are toxic to many microbes; however, these compounds may be degraded following a period of microbial adaption [74]. Biogeochemical parameters Measurements of various biogeochemical parameters such as dissolved oxygen (DO), redox potential, CO2, and other parameters such as NH4

+, NO3-, NO2

-, SO42-, S2-

and Fe2+ will give an indication of the existing (natural or intrinsic) microbial metabolic activity at the site [37]. Environmental constrains

Temperature Microbial metabolism is substantially affected by temperature [104]. Most microorganisms grow well in the range of 10 to 38°C. Technically it is extremely difficult to control the temperature of in-situ processes, and the temperature of ex-situ processes can only be moderately influenced, sometimes with great expense. Fortunately, although temperatures within the top 10 m of the subsurface may fluctuate seasonally, subsurface temperatures down to 100 m typically remain within 1° to 2°C of the mean annual surface temperature suggesting that bioremediation within the subsurface would occur more quickly in temperate climates [105]. pH The pH range within which most bioremediation processes operate most efficiently is approximately 5.5 to 8. It is no coincidence that this is also the optimum pH range for many heterotrophic bacteria, the major microorganisms in most bioremediation technologies. The optimum pH range for a particular situation, however, is site-specific. The pH is influenced by a complex relationship between organisms, contaminant chemistry, and physical and chemical properties of the local environment. Additionally, as biological processes proceed in the contaminated media, the pH may shift and therefore must be monitored regularly. The pH can be adjusted to the desired range by the addition of acidic or basic substances (i.e., mineral acids or limestone, respectively). However changes in soil pH will influence dissolution or precipitation of soil metals and may increase the mobility of hazardous materials. Therefore, the soil buffering capacity should be evaluated prior to application of amendments [29]. The effect of pH on permeability of soils and sediments is not fully understood but it seems that soil pH has also significant effect. Soils have a negative permanent charge and a pH-dependent variable charge. Therefore, pH affects soil dispersion and its permeability.


A typical volcanic ash soil has a large amount of pH-dependent charge. Its saturated hydraulic conductivity decreases under low and high pH conditions. When the predominant anion is sulphate, hydraulic conductivity does not decrease even at low pH. However, the saturated hydraulic conductivity of soils with montmorillonite and kaolinite at pH 9 is smaller than that at pH 6 [21]. Moisture content - water activity Moisture is a very important variable relative to bioremediation. Moisture content of soil affects the bioavailability of contaminants, the transfer of gases, the effective toxicity level of contaminants, the movement and growth stage of microorganisms, and species distribution. During bioremediation, if the water content is too high, it will be difficult for atmospheric oxygen to penetrate the soil, and this can be a factor of limiting growth efficiency and determine the types of organisms that can flourish. Various workers in the field have reported that the water content of the soil should be between 20 and 80%. In cases where no extra source of oxygen is being provided (for example, bioremediation of surface contamination), 20% moisture may be adequate; however, if a continuous recirculation system (pipe networks) is being used for deeper contamination, 80% water content would be more appropriate [74]. Soil moisture is frequently measured as a gravimetric percentage or reported as field capacity. Evaluating moisture by these methods provides little information on the “water availability” for microbial metabolism. Water availability is defined by biologists in terms of a parameter called water activity (aw). In simple terms, water activity is the ratio of the system’s vapour pressure to that of pure water (at the same temperature) [37, 74]. Redox potential The redox potential of the soil (oxidation-reduction potential, Eh) is directly related to the concentration of O2 in the gas and liquid phases. The O2 concentration is a function of the rate of gas exchange with the atmosphere, and the rate of respiration by soil microorganisms and plant roots. Respiration may deplete O2, lowering the redox potential and creating anaerobic (i.e., reducing) conditions. These conditions will restrict aerobic reactions and may promote anaerobic processes such as denitrification, sulfate reduction, and fermentation. Many reduced forms of polyvalent metal cations are more soluble (and thus more mobile) than their oxidized forms. Well-aerated soils have an Eh of about 0.8 to 0.4 V; moderately reduced soils are about 0.4 to 0.1 V; reduced soils measure about 0.1 to -0.1 V; and highly reduced soils are about 0.1 to -0.3 V. Redox potentials are difficult to be measured in the soil or groundwater and are not widely used in the field [29]. Mass transfer characteristics Mass transport characteristics are used to calculate potential rates of movement of liquids or gases through soil and include: Soil texture, unsaturated hydraulic conductivity, dispersivity, moisture content vs. soil moisture tension, bulk density, porosity, hydraulic conductivity and infiltration rate [40, 106-108]. Site hydrogeologic characteristics Hydrogeologic factors for consideration include aquifer type, hydraulic conductivity, hydrogeologic gradient, permeability, recharge capability, depth to groundwater, moisture


content/field capacity, thickness of the saturated zone, homogeneity, depth to contamination, extent of contamination, and plume stability. These are only some parameters that should be factored into the design of any bioremediation system [37, 40, 72, 106-108]. Biostimulation Biostimulation is the modification of the environment by adding nutrients, such as nitrogen and phosphorus, as well as oxygen and other electron acceptors to stimulate the rate of biological degradation of contaminants by indigenous microorganisms. This alternative is also chosen when a natural microbial population exists at the site which has the potential to degrade the chemicals, but is actually lacking oxygen, nitrogen or other nutrients to degrade them. The missing component(s) can then be introduced into the system and the degradative activity of the microbial community can be induced. Most bioremediation systems employ some form of biostimulation [37, 58]. Commonly used water soluble nutrients include mineral salts (e.g., KNO3, NaNO3, Ca(NO3)2, NH4NO3, (NH4)2SO4, K2HPO4, (NH4)2HPO4, MgNH4PO4,), anhydrous ammonia (NH3), urea (NH2)2CO and many commercial inorganic fertilizers (e.g., the 23:2 N:P garden fertilizer used in the Exxon Valdez case). There is some disagreement on the efficacy of various inorganic nitrogen salts for bioremediation. It is generally not agreeable which nitrogen source ammonium or nitrate is the most preferable. Water soluble nutrients are usually applied in the field through the spraying of nutrient solutions or spreading of dry granules. This approach has been effective in enhancing oil biodegradation in many field trials [109]. Compared to other types of nutrients, water soluble nutrients are more readily available and easier to manipulate. Another advantage of this type of nutrient over organic fertilizers is that the use of inorganic nutrients eliminates the possible competition of carbon sources. However, water soluble nutrients also have several potential disadvantages. First, they are more likely to be washed away by the actions of tides and waves when applied in seawater, and second, inorganic nutrients, ammonia in particular, should be added carefully to avoid reaching toxic levels [17]. Slow released or control released granular nutrient sources have been also used for bioremediation of oil spill contaminated areas, providing continuous release of nutrients over extended time period [110]. Slow release fertilizers are normally in solid forms that consist of inorganic nutrients coated with hydrophobic materials like paraffin or vegetable oils (e.g MaxBac® [44]). Other examples of slow released nutrients are sulfur-coated urea and Osmocote®, slowly soluble materials (such as metal ammonium phosphates), or materials that must be microbially mineralized to release nitrogen (e.g. organic fertilizers and urea formaldehydes). To maximize effectiveness, controlled-release nutrients should be released at a rate equivalent to microbial demand. However, the main disadvantage of these materials is that the nutrient release rate cannot be easily controlled in the field [109, 111-113]. Another approach to overcome the problem of water soluble nutrients being rapidly washed out is to utilize oleophilic organic nutrients (e.g. Inipol EAP22® - Elf Atochem of North America, Inc. - an oleophilic urea-based fertilizer) [44, 58, 114, 115]. Since oil biodegradation mainly occurs at the oil-water interface and since oleophilic fertilizers are able to adhere to oil and provide nutrients at the oil-water interface, enhanced biodegradation should result without the need to increase nutrient concentrations in the bulk pore water [114]. According to the vendor, Inipol EAP-22 has several advantages:


(a) Optimizes ratio between carbon, nitrogen, and phosphorus, (b) releases nutrients over time, (c) inhibits the formation of water-in-oil emulsions and (d) is completely biodegradable [44]. Some commercial proprietary biostimulation agents are branded by the names FyreZymeTM, (Ecology Technologies International, Inc.) and HCZyme, (CCharbon Consultants). Both are a combination of bacterial growth enhancing agents, extracellular enzymes, and surfactants. The bacterial growth enhancers increase natural biological processes by stimulating a logarithmic growth phase of indigenous microbes in soil and water, while extracellular enzymes initiate the oxidation that degrades petroleum-based contaminants, and surfactants help desorb the petroleum bound to soil particles. FyreZymeTM also contains a biodegradable compound that adds oxygen to the soil, thereby facilitating hydrocarbon degradation [44]. Biostimulation has been successfully applied in marine oil spills [18, 39, 58, 116] and polycyclic aromatic hydrocarbon (PAH)-contaminated soils [117]. A review of oil spill biodegradation literature is given by Swannell, Lee & McDonagh [109]. Bioaugmentation Bioaugmentation is the soil amendment with non-indigenous allochthonous microorganisms or cultivated indigenous species or engineered microbes via inoculation, or use of bioproducts, (e.g. enzymes i.e. lipases, proteases, cellulases etc.), to enhance the biodegradation of contaminants [58]. The introduced microorganisms augment, but do not replace, the resident microbial population. Usually bacteria with necessary catalytic activities and other required characteristics are injected directly into the polluted site usually together with nutrients. Bioaugmentation can be necessary also in cases where bacteria with the required catalytic activity although present at the site degrade the pollutants incompletely and/or at a very slow rate. Successful microbial inoculation requires a range of factors: (1) the population must be capable of surviving and growing in the new environment; (2) the microorganisms must retain their degradative abilities under the new conditions; (3) the organisms must come in contact with the contaminants; and (4) the electron donors/acceptors and nutrients necessary for microbial growth and contaminant degradation must be made available to the population [118]. Once the microorganisms are injected into the aquifer, there must be some mechanism for dispersing them throughout the biostimulation zone before they attach to the solid matrix and carry out the degradation reaction of interest. Cell transport within porous media is highly dependent on the characteristics of both the solid media and the microbial cells. Experiments have shown that the conditions that best promote microbial transport in porous media include (in order of their importance) highly permeable media, ground water of low ionic strength, and small-diameter cells [119]. Unfortunately the microorganisms that are efficient in the laboratory conditions do not cope well in the real world. Under natural conditions, laboratory strains face unfavourable nutritional and physicochemical conditions. They have to compete with already established indigenous communities and they have to withstand a variety of predators. Moreover, there is often a mismatch between the normal habitat of the introduced species and the ecological conditions in which they are placed. Finally, when biostimulation and bioaugmentation are used simultaneously, it is a common finding that the added nutrients favour mostly indigenous populations so that they overgrow the introduced species. The experience obtained from Exxon Valdez disaster at the coast of Alaska in 1989 in conjunction with research following other oil spills, demonstrated that in the case of


the Exxon Valdez clean-up, biostimulation of indigenous microbial consortia through the use of nitrogen and other fertilizers is more effective than bioaugmentation through seeding with exogenous organisms [112, 120, 121]. It seems that in most cases biostimulation with nutrients is often more effective concerning biodegradation rates than bioaugmentation with inoculums [3, 58, 116]. Most microbial inoculants or additives sold for use in bioaugmentation approaches have historically been blends or consortia of microorganisms, purportedly tailored for the types of compounds found in the target waste stream. Some commercial proprietary bioaugmentation products are branded by the names M-1000TM, (Micro-Bac International, Inc), Bac-Terra, (Microbe Technology Corporation), ABR (Augmented Bioreclamation) Microbial Blends, (SSybron Chemicals, Inc.), WST Bioblends, (Waste Stream Technology, Inc.), PetroKlenz, (Aqualogy BioRemedics), GT-1000, (Bio-Genesis Technologies). Information for these products is limited and originates mainly from the production/application companies. For example, Bac-Terra includes natural organic matter with a blend of microbial consortia capable of working in both aerobic and anaerobic environments, including psychrophilic, mesophilic, thermophilic, bacteria cultures for use at temperatures ranging from 28 to 240°C. Bac-Terra requires soil moisture greater than 20%. PetroKlenz is a dry powder containing specific cultured facultative anaerobes, naturally occurring microbes that were originally derived from soil and have been preserved through advanced drying techniques. The various strains are grown individually in pure culture and compounded together with powdered wetting agents, buffering agents, and other synergists that allow the organisms to readily adapt to the treatment environment. The organisms have been carefully matched to complement each other for the effective biodegradation of hydrocarbons. GT-1000 is a synergistic group of non-pathogenic microorganisms. According to the vendors most of these products have been used successfully in bench, pilot or full scale levels to demonstrate the ability to degrade pollutants such as benzene, ethylene, toluene, and xylene (BETX), volatile aromatic hydrocarbons, oil and grease, coal tars, phenolic compounds, and chlorinated organic solvents [44]. Other microbial bioaugmentation products are cited in the literature [20]. Up-to-date, there has been little convincing evidence for successful in situ remediation of aquifers resulting from introduced microbial populations. The limitation of distributing the exogenous microbial cultures in the subsurface and the question of long-term survivability of these lab-grown cultures under field conditions also discourage bioaugmentation. Bioaugmentation may play a prominent role in bioremediation when the release of genetically engineered organisms will be permitted [17, 20, 30, 37, 47]. Monitored natural attenuation Natural attenuation, often called intrinsic remediation, intrinsic bioremediation, bioattenuation, or monitored natural attenuation (MNA), consists of unassisted and unenhanced physical, chemical, and biological processes (e.g. biodegradation, abiotic transformations, mechanical dispersion, dilution, evaporation, volatilization and adsorption) that act to limit the migration and reduce the mass, toxicity, mobility, volume, or concentration of contaminants in soil, sediment, or groundwater [15, 19, 44]. Natural attenuation, as an in-situ technology, is very important because it is often technically infeasible to clean a contaminated site to regulatory cleanup levels for a variety of conditions including the presence of low-permeability soils, the inability to remove all the contaminants from the individual soil particles etc.


Natural attenuation is usually considered as the ‘baseline option’, and although it takes place without human intervention, the technology is not equal to a “do-nothing” or “no further action” approach. The difference between the use of “natural attenuation” and “no further action” as a remedial strategy is that natural attenuation requires thorough documentation and extensive monitoring of the role of microorganisms and other attenuation processes in eliminating the target contaminants. This implies that additional site characterization and development of a groundwater monitoring phase for an acceptable period of time may be necessary. Natural attenuation, if properly demonstrated, increases the overall protection of the environment by either containment or destruction of contaminants. No further action, on the other hand, implies that no additional investigation is required regardless of whether the contaminants of concern are degrading or migrating. Natural attenuation also serves as (1) an interim measure until future technologies are developed, (2) a managerial tool for reducing site risks, and (3) a bridge from active engineering (i.e., pump-and-treat, vapour extraction, etc.) to no further action. No further action, however, may be preferable to natural attenuation in certain instances. Very low risk situations may be better served since it eliminates the need of continued monitoring and further documentation. Sites with low levels of contaminants or nondiscernible plumes may be better candidates for no further action. Furthermore, very minor releases of hydrocarbons to the subsurface may not be sufficient to support bioremediation [72]. A site-specific, cost–benefit analysis is required to determine if an active remediation system or MNA would be the most effective remediation option. MNA may be an appropriate cleanup option when the facility can demonstrate that the remedy is capable of achieving specific ground-water cleanup levels in a reasonable cleanup time frame. If MNA is chosen then there are several costs associated with the implementation of it. These costs include modelling contaminant degradation rates to determine if natural attenuation is a feasible remedial alternative, subsurface sampling and sample analysis (potentially extensive) for determining the extent of contamination and confirming contaminant degradation rates and cleanup status. Regular operation and maintenance (O & M) costs are required for monitoring to verify degradation rates and maintain data on contaminant migration. In some cases, such long-term monitoring may be more expensive than active remediation [44]. When natural attenuation is permitted as a remediation strategy, extensive monitoring is required. It is beyond the scope of this paragraph to report extensive strategies for MNA protocols. Interested readers can find information in the literature [35, 40, 72, 122]. While natural attenuation will not be a suitable remedy for all contaminated sites, it does offer the potential advantages of in-situ technologies:

• Generates less secondary wastes, reduced risk of human exposure during treatment, reduced potential for cross-media transfer of contamination.

• Operates in-situ with minimal site disturbance. • Can be used in conjunction with other remediation technologies. • Reduced need for on-site structures associated with cleanup. • Potentially reduces overall remediation costs.

However, the potential limitations of natural attenuation include [15, 44]:

• It is well established as a remediation approach for only a few types of contaminants, (e.g. benzene, toluene, ethylbenzene, and xylene referred as BTEX, oxygenated hydrocarbons, low-molecular-weight alcohols, ketones, esters, and methylene chloride).


• Generally requires longer time frame for remediation, (for many years or decades). • Requires more involved site characterization and monitoring. • Toxicity and mobility of transformation products may be greater than that of the

parent compound. Some compounds can form hazardous by-products that in some cases can persist in the environment.

• Changes in environmental or site conditions may allow contaminant migration. • There is a potential for remobilization of previously stabilized metals and

radionuclides. • Public may see natural attenuation as a “do-nothing” approach. Biotransformation of metals, metalloids and radionuclides The toxicity and mobility of the elements depend primarily on their speciation, which is significantly influenced by soil pH, redox conditions, and surface chemistry. All of these are environmental factors that can be optimized by manipulating microbial activities to reduce the risk posed by heavy metals in aquifer environments. Microorganisms cannot convert metals to different elements. However, they can modify the microenvironment around the microbial cell and can catalyze oxidation, reduction, methylation and dealkylation reactions that affect the solubility and mobility of many metals. In addition, the microbial cell offers a large number of possible physico-chemical mechanisms of interaction (e.g. complexation, coordination, chelation, ion exchange, adsorption, microprecipitation) with soluble metal, radionuclide and metalloid species resulting in immobilization of them [24, 52, 123-128]. Figure 6, presents different possible metal mobilization/immobilization mechanisms.

Figure 6. Biological metal mobilization/immobilization mechanisms.


Bioprecipitation Metabolic mediated processes modify the environment around the microbial cell. Under aerobic conditions microorganisms grow by transferring the electrons available from the electron donor molecule (in the case of bioremediation usually an organic contaminant) to the oxygen dissolved or transferred from the ground to the aquifer. Organic carbon is mineralized to carbon dioxide and oxygen is reduced to water. The produced dissolved carbon dioxide increases the alkalinity and pH of the cells microenvironment and the excess bicarbonate favours the precipitation of metal ions as metal hydroxides Me(OH)x or carbonate Me2(CO3)x [103, 124]. In subsurface environment, where anaerobic conditions prevail, nitrates can act alternatively as terminal electron acceptors. The process known as denitrification is widely used in municipal wastewater treatment units. Removal of nitrates proceeds through nitrite intermediate, to nitrogen gas. From the oxidation of the carbon source, bicarbonates are also produced increasing the pH of the medium. Soluble metal ions are precipitated as metal hydroxides or carbonates with reactions similar to the case of aerobic metabolism. A wide range of bacteria such as Pseudomonas and Alcaligenes sp. immobilize dissolved metal species by the previously described action [103, 124, 129]. Finally under anoxic conditions and in the presence of sulphates, sulphate reducing bacteria (SRBs) can grow using sulphates as terminal electron acceptor. Genera such as Desulfovibrio, Desulfobacter, Desulfococcus, Desulfosarcina, Desulfomonas and Desulfotomaculum, Desulfomicrobium, Desulfobulbus can oxidize simple organic molecules such as ethanol, lactate or acetate to produce excess of sulphides in their microenvironment [102, 130]. Under such conditions metal ions can immobilize as metal sulphides, which are the most insoluble forms among the various metal precipitates (e.g. hydroxides, carbonates and/or phosphates) [103, 123, 124, 129-132]. Successful large scale ex-situ applications of SRBs have been documented in the literature [126]. Denitrifying microorganisms that use nitrate as an electron acceptor as well as sulphate-reducing bacteria which reduce sulphate to sulphide have been stimulated in situ by injecting acetate as a primary substrate and nitrate/sulphate as the electron acceptor in many bioremediation schemes [29, 103, 130]. Bioreduction – Biooxidation A wide variety of microorganisms can catalyze the reduction of heavy metals, such as Fe(III) to Fe(II), Mn(VI) to Mn(II), Cr(VI) to Cr(III), Se(VI) to Se(IV) or Se0, As(V) to As(III), Mo(VI) to Mo(IV) and U(VI) to U(IV). In such bioreduction processes reduced elements can serve as electron acceptors in alternative microbial respiration, or reduced by enzymes without energy production [125, 127]. Few examples for redox couples are presented below. Chromium appears in the environment in two oxidative states: Cr(VI) and Cr(III). Chromium(VI) is mobile in groundwater and is considered carcinogenic and more hazardous than chromium(III), which is a cation that tends to bind strongly to aquifer material [29, 133]. A wide range of microorganisms are capable of enzymatic reduction of Cr(VI) [134]. Some cells also obtain energy from Cr(VI) reduction, although most cells mediate this reaction cometabolically and do not harvest energy. For example, dissimilatory reduction of soluble Cr(VI) by Desulfovibrio desulfuricans, D. vulgaris is mediated by enzymatic reactions [135]. Soluble enzymes are thought to be responsible for the reduction of chromate by Bacillus sp., Pseudomonas sp., and E. Coli [136].


Aerobic reduction is thought to be a detoxification process where cells use a soluble enzyme to reduce Cr(VI) to Cr(III) internal or external to the plasma membrane. Reduction of Cr(VI) may also proceed through the use of CrO4

2- as a terminal electron acceptor during anaerobic respiration [124, 137]. The produced Cr(III) can be considered less mobile as it can be precipitated in the form of insoluble Cr(OH)3. However, experimental results have shown that in the presence of many organic molecules Cr(III) remains mobile forming stable soluble metal organic complexes [137-139]. Once reduced, Cr(III) cannot be oxidized by microbes back to Cr(VI). When considering natural attenuation of Cr(VI), care should be taken that there are no oxidized forms of manganese in the aquifer matrix (e.g., MnO2). Such minerals are known to abiotically oxidize and remobilize Cr(III) back to Cr(VI) [15, 127]. Microbial reduction of soluble U(VI) could be an attractive alternative for in situ bioremediation of uranium-contaminated groundwater [140-145]. It has been demonstrated that soluble U(VI) in the form of uranyl ions UO2

2+ can be reduced to insoluble U(IV) by microorganisms belonging to the genera Geobacter, Shewanella as well as Desulfovibrio desulfuricans, Pseudomonas fluorescens, and Deinococcus radiodurans [143-145]. U(IV) is finally immobilized as black UO2(s) precipitate [146, 147]. However the process is sensitive to nitrate concentration in the ground water, because nitrate compete uranyl ions in the bioreduction process and affects the stability of U(IV) precipitate [133, 148, 149]. Enzymatic reduction of Mo(VI) (as molybdate, MoO4

2-) to Mo(IV) by D. desulfuricans with both lactate and hydrogen as electron donors has been also reported [141, 150]. Mo(VI) reduction in the presence of sulphide results in the extracellular precipitation of the black mineral phase, molybdenite MoS2(s). The most important transformation process in the environmental fate of mercury is biotransformation. Any Hg form entering sediments, groundwater, or surface water under the appropriate conditions can be microbially converted to the methylmercuric ion. Sulfur-reducing bacteria are responsible for most Hg methylation in the environment, with anaerobic conditions favouring their activity. The methylation of elemental Hg plays a key role in environmental cycling of Hg. Methylated Hg, the most common Hg form, is mobile and readily taken up by organisms including some higher plants. Humic substances are known to mediate the chemical methylation of inorganic Hg by releasing labile methyl groups [29]. Mercury reduction may be a mechanism for detoxification of media containing mercury. Reduction of Hg2+ to elemental mercury occurs quite readily and is enhanced by bacterial enzymes [92]. Energy is probably not obtained by use of Hg2+ as an electron acceptor [124, 133]. Oxyanions of selenium (SeO4

2-, SeO32-) can be used in microbial anaerobic

respiration as terminal electron acceptors providing energy for growth and metabolism. Their reduction can be coupled to a variety of organic substrates, e.g., lactate, acetate and aromatics, with the bacteria found in a range of habitats and not confined to any specific genus. These organisms, and perhaps even the enzymes themselves, may have applications for bioremediation of selenium contaminated environments. Microbial reduction of the soluble oxidized form of selenium, Se(VI), to insoluble elemental selenium, Se0, is possible by microorganisms that conserve energy to support growth from Se(VI) reduction [141, 151, 152]. This natural mechanism can be used for the removal of selenium from contaminated surface and groundwater. This metabolism may be employed in the environment for selenium bioremediation. Details on the biogeomicrobiology of selenium can be found in the literature [29, 127, 133, 153, 154].


Bioreduction of elements does not always reduce their mobility and could be undesirable. For example, under anaerobic conditions, As(V) in the form of arsenate (AsO4

3-) may also serve as an alternate electron acceptor and be reduced to more toxic and more mobile As(III) in the form of arsenite (AsO2

-). This process is reversible because arsenite can be reoxidized to arsenate and thus re-immobilized under aerobic conditions. Similarly, Mn(IV) is reduced to the more toxic and more mobile Mn(II) under anaerobic conditions, with a reaction which is also reversible [15]. Details on the biogeomicrobiology of arsenic can be found in the literature [29, 127, 133, 155-158]. Biosorption Biosorption can be defined as the selective sequestering of metal/metalloid or radionuclide soluble species by microbial cells that results in immobilization of them. Metal sequestering by different parts of the metabolically active or inactive cells can occur via various mechanisms: complexation, chelation, coordination, ion exchange, precipitation, reduction [159, 160]. Biosorption is a process with some unique characteristics. It can effectively sequester dissolved metals from very dilute complex solutions with high efficiency. This makes biosorption an ideal candidate for the treatment of high volume low concentration complex waste-waters. However, today biosorption is not considered as a competitive stand alone technology because the industrial applicability of the process is rather limited and pilot applications have shown the limitations associated with the use of inactive microbial biomass mainly due to the cost of formulating it into an appropriate biosorbent material. Furthermore, high concentrations of co-ions present in the treatment water have negative effect on the uptake of the targeted metals by the immobilized microbial biomass, and the reduced resilience of the biological material, made recycling and reuse of the biosorbent even more difficult. However in the cases of metabolically active microbial cells, biosorption contributes in the overall sequestering and immobilization of metal ions as a parallel mechanism together with other metabolically mediated mechanisms such as bioprecipitation and bioreduction [159-163]. Phytoremediation Phytoremediation is defined as the engineered use of vegetation to contain, sequester, extract, accumulate, remove, degrade and/or detoxify inorganic and organic contaminants from soils, sediments, surface waters, and groundwater [29, 53]. Plant-based remediation systems are generally considered passive, low-cost, low-technology processes and employ common plants including trees, vegetable crops, grasses, and even annual weeds to treat heavy metals, inorganic ions, radioactive elements, and organic compounds. Phytoremediation is considered as an alternative bioremediation option appropriate for soils having properties that impede the success of conventional technologies (e.g., low permeability, saturation, dense structure, mixtures of contaminants). When the appropriate plants are cultivated in contaminated soils the root system functions as a dispersed uptake system. Contaminants are taken up with soil water and degraded, metabolized, and/or sequestered in the plant, while evapotranspiration from aerial parts maximizes the movement of soil water through the plant. Certain plants have been identified that can take up and concentrate metals and other inorganic molecules from soil into leaves, stalks, seeds, and roots. Commonly used hyperaccumulators include sunflower (Helianthus agnus), Indian mustard (Brassica juncea), crucifers (Thlaspi


caerulescens, T. elegans), violets (Viola calaminaria), serpentines (Alyssum bertolonii), corn, nettles, and dandelion, [15]. The cultivated plants subsequently can be harvested and treated by incineration, composting, or anaerobic digestion to concentrate and/or recover the pollutants [29]. Phytoremediation technology can be divided into of two main broad categories depending upon whether the removal of contaminants or stabilisation of geochemical conditions in the soil is accomplished. Many different mechanisms for pollutants uptake removal or stabilization have been identified [8, 15, 92]: In phytodegredation, organic pollutants are converted by internal or secreted enzymes into compounds with reduced toxicity. Like phytodegradation, rhizosphere degradation or rhizodegradation involves the enzymatic breakdown of organic pollutants, but through microbial enzymatic activity. These breakdown products are either volatilized or incorporated into the microorganisms and soil matrix of the rhizosphere. The types of plants growing in the contaminated area influence the amount, diversity, and activity of microbial populations thus there is a direct independence between the planted vegetation and the dominant microbial species in the soil matrix of the rhizosphere. Phytoextraction involves the removal of toxins, especially heavy metals (e.g. Cd, Ni, Hg) metalloids (e.g Se) and radionuclides, by the roots of the plants with subsequent transport to aerial plant organs. Plants can also remove toxic substances, such as organics, from the soil through phytovolatization. In this process, the soluble contaminants are taken up with water by the roots, transported to the leaves, and volatized into the atmosphere through the stomata. Rhizofiltration removes contaminants from water and aqueous waste streams, such as agricultural runoff, industrial discharges, and nuclear material processing wastes. Absorption and adsorption by plant roots play a key role in this mechanism, and consequently large root surface areas are usually required. Finally, phytostabilization can be used to reduce the erosion, leaching and mobilization of soil contaminants which result in aerial or waterborne pollution of additional sites. In phytostabilization, accumulation by plant roots or precipitation in the soil by root exudates immobilizes and reduces the availability of soil contaminants. Plants growing on polluted sites also stabilize the soil and can serve as a groundcover thereby reducing wind and water erosion and direct contact of the contaminants with animals. An example of a simple phytoremediation system in use for many years is the constructed wetland, in which aquatic plants such as water hyacinths are cultivated to remove contaminants (metals, nitrate, etc.) from municipal or industrial wastewater [29]. Phytoremediation has been proven useful for soils contaminated with relatively immobile contaminants to shallow depths and for the case of organic pollutants is generally applicable for moderately hydrophobic material. Examples of these are toluene, benzene, PAHs, xylenes, ethylbenzene and many chlorinated solvents. Organic compounds can be degraded or immobilized in the root zone or incorporated into shoot tissues and metabolized. The advantages of phytoremediation technology are summarized below [44]: • Operates in-situ and is solar driven. • Costs are approximately 20 to 30% of costs associated with mechanical treatments. • It has high public acceptance. • It is applicable to many remediation scenarios including large contaminated surface areas.

However, some potential disadvantages associated with phytoremediation/plant-assisted remediation techniques have been identified [44, 55, 61]:


• Treatment is generally limited to shallow soils within 90 cm from the surface and groundwater within 3 m from the surface.

• High concentrations of hazardous materials can be toxic to plants. • It involves the same mass transfer limitations as other biotreatment technologies. • Treatment time is relatively long (usually more than one growing season). • It may be seasonal, depending on location. • Climatic or hydrologic conditions may restrict the growth rate of certain plants. • Contaminants may enter the food chain via animals (herbivores) or insects that

consume plant material containing contaminants. • Degradation products may be mobilized into ground water or bioaccumulated in

animals. • It can transfer contamination across media, e.g., from soil to air. • It is not effective for strongly sorbed (e.g., PCBs) and weakly sorbed contaminants. • Disposal of secondary waste arising from the harvest of plants is problematic. • The toxicity and bioavailability of biodegradation products is not always known. • It is still in the demonstration stage. • It is unfamiliar to regulators.

Concerning cost aspects, phytoremediation utilizes solar energy, thus it requires less energy inputs. This factor reduces operating costs. In addition as phytoremediation is slower treatment process and last longer, expenses are also spread out over a greater time period than other technologies. The result is lower annual costs. However, phytoremediation does have several cost components that are unique to the technology. These components often include the following [44]:

• Plant or tree stock and/or seeds • Fertilizers, pesticides, and additional soil amendments • Agricultural equipment for amendments application, tilling, and/or harvesting • Irrigation equipment and a water source • Tubes for stimulating deep root growth (collars) • Pest control devices • Supplies, equipment, and/or analyses for testing plant tissues and environmental

conditions • Flow control devices • Plant litter collection, maintenance, pruning, mowing, and/or harvesting • Disposal of plant wastes The literature cited in this chapter gives more details concerning phytoremediation technology [11, 29, 35], phytoremediation of metals and radionuclides [9, 29], phytoremediation cost data as well as comparative economic values between phytodegradation and a conventional pump-and-treat method [44]. Conclusions Bioremediation is a multidisciplinary technology and successful application requires deep understanding of all the relevant scientific fields and attenuation processes. It seems that nowadays we have entered in the most interesting and intense phase of process development. Potentials and limitations of the technology are well documented in many


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