16

Soil Microbes: Optimizers of Soil System Sustainability and Reparation of Damaged Soils

Soil management, as well as the changes that naturally impact soil function with evolution of a soil‐based ecosystems, also affect systems associated with atmospheric process down to the more limited terrain of a backyard garden. An excellent example of this keystone relationship is exemplified by the large variety of microbes and higher plants affecting the size of soil aggregates and the rate of their formation. As introduced in Chapter 1, soil aggregation is a process that affects a wide variety of soil properties, such as, influx and efflux of soil gases, including those associated with the atmospheric processes leading to climate change (e.g., carbon dioxide, methane, nitrous oxide, and nitrogen oxides.). Burke et al. (2012) provide an example of how the variety of microbial functional groups can impact the fate of greenhouse gases in soil. Additionally, examples of the complexity and diversity of these interactions between plant community and soil minerals and management are discussed by Poirier et al. (2018). Indeed, soil microbes are major contributors to ecosystem function in general. As stated by Comerford et al. (2013)

“At this point of our examination of occurrence and functions of microbes in soil, one could easily reach the conclusion that the primary role of microbes in soils involves optimization of food and fiber yields, as revealed with both laboratory and field study. Such research has and will always contribute to our understanding of the importance of maintaining a healthy soil microbial community existence in a soil system but, in reality the soil microbes play a significant role in supporting all life on our planet. The all‐encompassing importance of soil mineral‐microbial interactions is a major support for all aspects of survival and sustainability of soil systems, be they for cropping, industrial usage, or are simply in jeopardy due to basic human societal activities.”

Thus, with these considerations of the role of microbes in bioremediation of all soils impacted by society in a negative manner, this chapter is presented as a capstone analysis of the importance of management of the microbial community for maximizing the productive and sustainability of soil systems in general and damaged soil systems specifically. It would perhaps be useful to consider the question “What are the types of systems that should be considered in our analysis of the types of human impacted ecosystems that should be considered in this analysis? Major human societal impacts soil organic matter can easily be traced to the times of the agricultural revolution (See Chapter 1). Many examples of situations where the levels of soil organic matter were reduced sufficiently to impede crop production—even a new, different type of ecosystem. Today, it is recognized that agricultural development and the subsequent loss of soil organic matter has resulted in the realization that cropping of soils is a major source of atmospheric carbon dioxide, therefor making soil a significant player in climate change. Optimization of best management plans for soils is a critical factor in managing of atmospheric carbon dioxide levels.

A soil ecosystem requiring bioremediation can be conceptualized in part as one in which the existing biological community has failed to function as might be anticipated. In a non‐disturbed system, biologically decomposable substances entering the soil are mineralized. The concentration of the residual organic material is the difference between the amendment and decomposition rates. In a chronic situation, it is reasonable to assume that some detectable equilibrium concentration of the soil organic amendment will be achieved. If the input ceases or the material is added at long time intervals, it is equally reasonable to anticipate that the added substance could be totally eliminated periodically. In contrast, in our hypothetically polluted soil, a toxic substance has accumulated to concentrations that inhibit and/or interfere with the normal function of the ecosystem; that is, the system failed. Note that this statement does not exempt anthropogenic management failure from consideration. People are part of and major influences on the ecosystem. Thus, the failure of the microbial community to rid its environment of a toxic organic material could be a result of its being overwhelmed by mismanagement of the soil.

Considering that the pollutant accumulated to toxic levels, we can conclude that normal soil processes are inadequate to ameliorate the situation. The inability of the soil microbes to prevent the accumulation of the organic pollutant could be due to a number of factors:

· No individual or groups of microbial species exist in the soil capable of mineralizing or detoxifying the contaminant.

· The rate of input of the pollutant may be greater than its decomposition rate, resulting in ecosystem degradation.

· Chemical, physical, or biological limitations of the decomposers may exist or may be created by the contamination of the soil system thereby preventing microbial removal of the pollutant (e.g., anoxic conditions may develop preventing occurrence of obligatory aerobic processes).

· The pollutant(s) were amended to the soil at levels that are toxic to the decomposers themselves.

· Conditions may be optimal for decomposition to occur, but the decomposer population could be physically separated from the decomposable pollutant. That is, the organic compound may be physically or chemically sequestered.

· Decomposition of the pollutant may create conditions in the soil wherein further mineralization is precluded (e.g., acidification of the system, depletion of molecular oxygen).

The objective of bioremediation is to manage soil physical, chemical, and biological properties in a manner that overcomes these limitations to biological activity and optimizes the capability of the microbial community to detoxify the soil.

The scientific basis for the management of the soil microbial community for remediation is built on the same biological and chemical foundations as occur in non‐polluted soil systems. This commonality between functional native systems and those with excessive accumulations of toxic organic chemicals is exemplified by considering the processes involved with organic matter accumulation in soil. Amendment of any soil with plant biomass fuels the activity of the biological community and results in the production of microbial biomass and humus. With a normally functioning soil system, this accumulation of organic matter has a positive influence on ecosystem stability. (Although of course, there is a limit to the amount of organic matter that can be accumulated in soil.) Incorporation of plant biomass components into soil humus results, in part, from covalent linkages with soil humic substances. The organic components may also be occluded from decomposers by diffusing into soil micropores, by non‐covalent stabilization within the three‐dimensional structure of humic acids, or by sorption onto soil minerals. Since these are time dependent processes, a slowly mineralized substance has greater chance of becoming sequestered or sorbed onto or within clay or humic substances than do more biodegradation susceptible compounds. Thus, an easily metabolized organic compound, such as glucose or a protein, is normally fully mineralized by the soil microbial community in a short time, whereas cellulose and lignin of plant biomass are commonly stabilized within the soil structure. Also, small pieces of intact plant tissue can be protected from microbial attack by becoming encased in soil aggregates. In this situation, the sequestration results from physical adhesion of the soil particulate to small pieces of plant material. From the view of the normally functioning soil and non‐toxic organic substances, these organic matter stabilizing processes result in enhancement of the soil system or an improved soil quality. In contrast, similar protection of toxicants can result in ecosystem degradation and a reduction of ecosystem or environmental health. The processes are the same, but divergent consequences to the ecosystem function occur.

In the negative case, disruption of soil aggregate structure can actually become a positive management solution. Encouragement of accelerated decomposition of preexisting toxicants may result in not only lower pollutant levels but also in positive aspects of soil structure (such as, disruption of existing soil aggregation levels). In this situation, the availability of the pollutant molecules can be increased via liberation of substances previously protected within soil aggregates. In the latter situation, reduction in or destruction of the soil aggregate structure could be considered in the management plan initially to liberate the hidden pollutant followed by management of the soil to improve soil structure. The latter objective could be accelerated by introduction of plant communities. Note that this analysis primarily involves studies of remediation of surface soils.

Contamination of a soil ecosystem with biodegradable organic substances creates an unique situation for the microbes involved with their decomposition as well as for the site manager charged with evaluating the potential for bioremediation and the associated ecosystem risk. Thus, this chapter is presented with the overall objective of providing a foundation for considering the biological, chemical, and physical principles underlying or controlling the potential for bioremediation of soil ecosystem and the ultimate limitations or risks associated with such remediated systems. The soil amendment can range from simple organics (singly or in complex mixtures, biochar and biochar amended fertilizer, or plant nutrients in complex mixtures. (for examples of these materials used in the field, see Cartmill et al., 2014; Chen et al. 2014; Gandarillas et al., 2019; Lin et al. (2008); Zhou et al. 2019). This analysis will necessarily commence with a consideration of the microbes involved, but the primary emphasis will involve consideration of the physical and chemical limitations of the processes as they relate to bioavailability of the contaminant. Note that the topics analyzed herein tend toward analysis of basic concepts of microbial bioremediation. Basically, this chapter will extend upon the analyses presented in the proceeding chapters with a goal of providing a foundation for a more detailed examination of the bioremediation principles as presented specifically in the bioremediation literature.

16.1 Foundational Concepts of Bioremediation

16.1.1 Bioremediation Defined

Soil bioremediation is defined as the management or utilization of the soil biological community to detoxify, immobilize or mineralize organic contaminants in the soil ecosystem. Specifically, this process ideally involves conversion of the pollutant totally to mineral forms (e.g., carbon dioxide, methane, ammonium, hydrogen sulfide, etc.), its conversion to forms with no significant negative health traits (generally viewed from a human health viewpoint, but this discussion includes the concept of “health” of the total biological community), and/or immobilization via covalent linkage to soil humic acids (humification). The biological community generally managed for bioremediation is the native soil microbial community, but higher plants (phytoremediation) may also be manipulated to enhance toxicant amelioration, especially for remediation of metal‐contaminated soils. The primary emphasis of this chapter is management of soil microbes (including those of the rhizosphere) to optimize reduction of the risk associated with an organic toxicant in soil. These microbes may be indigenous to the contaminated soil system or amended to the soil during the remediation process. That is, the metabolic capabilities of the indigenous microbial community may be enhanced with genetically engineered (gems) or other laboratory derived or cultured microbial strains. Site management includes manipulation of soil chemical and physical properties to optimize both bioavailability of the pollutant and the microbial capacity to transform it. Frequently bioavailability is discussed in a limited sense—that is, in relationship to human health or risk. For this discussion, bioavailability will be considered the availability of the organic pollutant to the totality of the biological community, especially the soil microbes involved in its detoxification and/or mineralization.

16.1.2 Conceptual Unity of Bioremediation Science

The foundation for our understanding of the processes involved in bioremediation is derived from a unity of biochemical‐microbiological‐ecological principles. These supporting principles for bioremediation are as follows:

· All relationships of comparative biochemistry apply equally in the axenic culture of microorganisms and in the field soil. A compound mineralized by one or a combination of pathways or processes in culture can be anticipated to be similarly transformed in a field soil. A caveat to this rule is that the rate and extent of the transformation is controlled by the environmental situation of the microbes involved, including stresses imposed on the microbes through their interactions with other members of the biological community. Thus, the reactions may proceed at faster or slower rates and different intermediates may accumulate in the field due to limiting conditions (nutrient, physical, chemical or other). Only a small proportion of the soil microbes have been cultured in the laboratory and their biochemical capacities characterized. Variations in many biochemical processes that have yet to be described may exist.

· All microbial growth requirements are the same whether in the laboratory culture and in the field. If a growth factor is required in culture for a specific enzyme to function, it must also be supplied in the soil system. Similarly, the microbes and enzymes are controlled by the same physical limitations (e.g., pH, temperature) in the field and culture. The microbes may be functioning at the extremes of their growth margin—hence a slower process rate is seen—but, the organisms will not function beyond their normal range without some environmental compensation (a phenomenon also observed in axenic culture (See Chapter 5). For example, provision by root exudates or by other members of the microbial community of a cofactor whose de novo synthesis is precluded by high temperature may allow the microbe to function at a higher temperature.

· All limitations resulting from ecological interactions apply, including the need to accommodate the presence of other microbes as well as adaptation to the physical and chemical properties of the microsite wherein the microbe lives. This principle has most likely been the most overlooked when developing genetically modified strains or when selecting variant strains of microbes for soil inoculation. A microbe amended to a soil ecosystem with the intent of stimulation of bioremediation processes must not only possess the requisite genetic information, but be capable of expressing that capability in situ. Additionally it must also have the capacity to become, at least temporarily, a part of the overall soil microbial community. That is, the alien microbe must be able to express the requisite biochemical activity in the community, become established in that community, and accommodate or adjust to all means of interactions (ammensulism, competition, etc.) with its neighbors—at least until its presence is no long required for bioremediation.

These underlying principles of bioremediation processes underscore the fact that successful management of the soil microbial community to ameliorate contamination with organic pollutants is truly an interdisciplinary endeavor. All involved must have an appreciation of the basic concepts of soil science as well as the principles of soil microbiology, microbial physiology, and ecology. This interdisciplinary aspect of bioremediation technology is further accentuated when the impact of environmental engineers in applying the principles in the field is considered.

16.1.3 Complexity of Remediation Questions

Bioremediation involves mitigation of chemical contamination of soils and thereby a reduction of the impact of the presence of the pollutants on ecosystem function. By the nature of the vast array of organic substances and of the diversity in the extent of their encounter with soil systems, the soil system managed for bioremediation can range from a few square meters to many hectares in area. The degree of intervention may range from the simple working of a compacted soil surface and amendment with fertilizer to the implementation of complex, engineered systems for pollutant retention and soil management developed to meet the needs for remediation of highly contaminated industrial sites. Definition of the breadth of chemicals contaminating soils extends far beyond the purview of this chapter, but the range in structural complexity can be gained from such reviews as that of Swobada‐Colberg (1995) on xenobiotic chemicals.

Implications of the importance of bioremediation extend across not only those aspects associated with quality of the specific soil of interest, but also encompass more general concerns of environmental or ecosystem health. First concerns with bioremediation must involve mineralization of the organic substances of interest or immobilization of the inorganic substance, as is involved with phytoremediation of metal contaminated soils, but these initial endeavors must be executed with considerations of the effect of system management on the general health of the total environment in mind. That is, management of the decomposer population has implications far beyond inducing a simple reduction in soil toxicant loading. For example, utilization of mycorrhizal fungi as mineralizers of organic contaminants also alters nutrient and water dynamics of the associated plant community. This enhancement of the mycorrhizal association will result in a stimulation of plant biomass productivity, both aboveground and belowground. This increased plant growth will result in augmented activity of the rhizosphere community and an improved soil aggregate structure (see Chapter 1), primarily due to the stimulation of biological activity via exudate production.

A perhaps expected effect of the increased plant growth relates to soil water dynamics. In a soil system without a plant community, the direction of movement of the soil water, and any solutes (contaminants) contained therein, would be downward. With the increased water use of the plant in the system, a significant portion of the soil water is drawn to the root, thereby decreasing the quantity lost through percolation beyond the root zone. Thus, less pollutant is lost from the system by leaching to ground water and greater quantities are retained within the biologically activity zone of the soil (See Clothier and Green (1997) for a review of the impact of roots on soil water dynamics).

16.2 The Microbiology of Bioremediation

Consideration of the principles of bioremediation can be divided into two primary areas: a) Properties of the microbes themselves affecting their interaction with and decomposition of the pollutant and b) Properties of the ecosystem controlling the bioavailability of the toxicant. Successful biological remediation of contaminated soils relies not only on the metabolic capability of the microbes but also on the probability of a favorable encounter between the toxicant and it decomposer community. That is, success depends on the bioavailability of the pollutant. Biodegradable organic substances entering a soil ecosystem can be mineralized at a reasonable rate (assuming no other limitation to microbial activity beyond those fulfilled by the organic amendment) as long as the pollutant is present within the soil interstitial water and existent in appropriate concentrations. Failure to fulfill the latter requirement can result in concentrations below that which allows efficient interactions with degradative enzymes as well as higher concentrations, which may be toxic to the microbe. Organic substances tend to have maximal concentrations in the available soil water immediately after entering the soil environment. During this time period, for metabolizable substances, the most probable fate of the organic substance is mineralization and incorporation of the carbon or metabolites into microbial biomass. A portion of the organic substance may be humified, sorbed onto soil minerals, leached through the profile, or volatilized. For bioremediation situations, this seemingly simple situation is made more complex by the fact that the distribution of the organic substrate becomes more complex with time, that is the proportion which becomes biologically unavailable increases with time in the soil system. Thus, for soil microbes, native or otherwise, to decompose soil contaminants, they must posses the genetic capability necessary to synthesize the requisite enzymes, be in an environment conducive to expression of their genetic potential, and have the capability to overcome the physical limitations of accessibility of the contaminant. The physical and chemical requirements for successful bioremediation of a soil site are the topic of the next section. This discussion will be limited to the properties of the microbial community that are requisite for successful biologically based remediation of chemically contaminated soil ecosystems.

16.2.1 Microbes as Soil Remediators

Soil microbes responsible for the decomposition of organic pollutants exist in three states:

· The decomposer populations are sufficiently high that immediate decomposition of the pollutant can occur. If the quantity of the pollutant soluble in the soil water is non‐toxic and available at concentrations sufficient for effective catalysis, then degradation would be anticipated to occur—assuming that no other limitations (nutrients, oxygen, etc.) exit.

· Microbes capable of decomposing the contaminant are present in the soil system in low numbers. The pollutant may induce development of sufficient populations to allow the ultimate degradation of the pollutant. Major factors of consideration in this situation include the time necessary for the population to develop (See Chapter 4 on microbial growth kinetics in soil) and the potential that other soil factors may limit the development of adequate microbial populations for biological remediation to occur.

· No indigenous organisms capable of decomposing the contaminant are present in the microbial community. In this situation, the potential to amend the soil with decomposer populations capable of surviving and functioning therein must be considered. This process may involve use of genetically modified organisms or simply adding soil or other amendment that contains the requisite populations. This process imposes an added limitation discussed above—the necessity of the alien microbes to overcome competitive interactions with the indigenous microbes. The existence of an “underutilized” carbon and energy source, the pollutant, in the system could provide an advantage to the amended or invading microbes.

If the appropriate organisms exist within the soil system for biological decomposition of the substance to occur, then there are several alternative processes that should be evaluated when assessing the potential for bioremediation. The microbes may metabolize the organic pollutant directly, with the product ultimately being mineralized through common intermediary metabolism pathways. The pollutant may be metabolized indirectly. In this situation, there may be no organisms present in the system capable of mineralizing the substance, but the contaminant may be transformed cometabolically and ultimately decomposed through the interaction of several members of the microbial community (a consortium). Further variation in the metabolic processes associated with transformation of the pollutant results from the microbial capability to utilize a variety of terminal electron acceptors. Thus, although aerobic decomposition of organic substances is generally initially considered in ecosystems contaminated with toxic substances, the capacity of bacteria to use alternative terminal electron acceptors, such as nitrate and sulfate, can be exploited to purify the soil. Each of these alternatives will be discussed in greater detail when the pathways for bioremediation are considered below.

16.2.2 Substrate‐Decomposer Interactions

16.2.2.1 Physiological Pathways for Bioremediation

Description of specific pathways for biodegradation is beyond the scope of this chapter. (See chapter 4 for a general discussion of microbial metabolism in soil.) Examples of the complexity or variety of alternative pathways through which soil organic contaminants can be decomposed are provided by (Young and Cerniglia, 1995). Perhaps, the biggest mental obstacle to maintaining a clean soil environment in the early industrial era was that generally held, rather optimistic concept of microbial infallibility (See chapter 4) combined with the confidence derived from the great knowledge that had been accumulated regarding the complexity of intermediary metabolism of soil bacteria and fungi. A portion of that trust in the capabilities of the soil microbes and the breadth of our understanding of their metabolic prowess was not misplaced. A vast array of biologically synthesized organic chemicals and industrially produced organic chemicals (xenobiotic chemicals) with structures that mimic the biologically synthesized substances are metabolized in soil. Of course, the full function of the microbial community is predicated on the precept that the proper environmental conditions must be maintained for optimal function of the soil community and that the pollutants do not enter the system at rates that require the microbial community to decompose them at rates faster than their maximum potential. More problematic are those xenobiotic chemicals that do not have structures resembling biologically synthesized molecules. These substances may be totally recalcitrant, metabolized in part (DDT conversion to DDE), humified (if sufficiently reactive chemically), or lost from the system through volatilization or leaching. Also, some of these chemcials may be slowly transformed cometabolically. For example, trichloroethylene (TCE) is decomposed cometabolically by microbes using methane or other simple carbon compounds as their carbon and energy source (See Wackett (1995) for a review of this topic.). Similarly, halogenated organic compounds can be dehalogenated cometabolically and the resultant product mineralized. A common example of this process is microbial reductive dechlorination of polychlorinated biphenyls (PCBs) (e.g., Bedard and Quensen, 1995). Approximately 1.4 billion pounds of PCBs have been released and have accumulated in the environment. In anaerobic environments the chlorine molecules can be removed from some of these substances thereby reducing the toxic risk of the contamination.

For many structurally complex organic chemicals, degradation relies on the existence of consortia or microbial assemblages to catalyze the processes. Microbial consortia are groups of microbial species that when acting together are capable of mineralizing a substance that no single member of the consortium can achieve alone. Perhaps the most commonly cited natural example of a group of members of the soil microbial community interacting to achieve complete mineralization of a complex substrate is the conversion of cellulose to methane. No single microbe is capable of mineralizing cellulose to methane; yet, in anoxic systems, methane production from cellulose is common. A variety of anaerobic bacteria (e.g., Clostridium spp.) convert the cellulose to the simple organic acids that are then utilized by the methanogens in their production of methane. Similar interactions are easily envisioned to occur in the decomposition of xenobiotic substances. The dehalogenation of PCBs described above can be seen as an initial step in a complex decomposition pathway of a seemingly environmentally stable substance. A scenario can be described wherein the PCBs are dehalogenated under anaerobic conditions. The products then diffuse or are transported to an aerobic site where the biphenyl can be mineralized.

The best examples of exploitation of dehalogenation reactions for remediation of organic wastes are derived from studies of contaminated sediments, but there is no reason to anticipate that under comparable conditions such reactions would not occur in a soil ecosystem. A particularly interesting study is one in which terminal electron acceptors other than molecular oxygen were used for the degradation of monochlorobenzoate isomers by Nile river sediments (Kazumi et al., 1995). In this study, the monochlorobenzoates were degraded under denitrifying, methanogenic, iron reducing, and sulfidogenic conditions. Conversely, a study by Fliermans et al. (1988) demonstrated the capacity of aerobic microbial consortia to convert trichloroethylene (TCE) to hydrochloric acid and carbon dioxide. In the latter example, the consortia were isolated from subsurface sediments and grown in enrichment cultures. Other examples of microbial consortia based mineralization include decomposition of dicyclopentadiene (Stehmeier et al., 1996) and a complex mixture of aromatic compounds contained in olive oil mill effluent (Zouari and Ellouz, 1996). Similarly, in a complex biofilm, Field et al. (1995) demonstrated decomposition of aromatic pollutants in consortia of aerobic and anaerobic bacteria. Also, the anaerobic degradation of pentachlorophenol can be catalyzed by a methanogenic consortium (Juteau et al., 1995).

These examples demonstrate the variety of organic chemicals decomposed by microbial consortia and underscore the importance of the entire soil microbial community in bioremediation. This community or mixed culture per se involvement in bioremediation is particularly significant when it is considered that most situations wherein the indigenous soil microbial community is utilized to purify the systems, a mixture—generally complex—of organic compounds is encountered. Soils are rarely contaminated with a pure xenobiotic chemical. Even pesticides, which are commonly conceptually spoken of as if a single chemical substance were reaching soil, are commonly added to the soil‐plant system in petroleum or other appropriate carrier mixture to aid dispersion and to facilitate dilution of the active ingredient to the appropriate dosage.

16.2.2.2 Role of Biofilms in Soil Remediation

This admixture of a variety of potential carbon and energy sources encourages development of interactive soil microbial communities. This observation leads to the question of the form or nature of the physical association of these microbes within the soil environment. As was indicated in chapter 2, bacteria exist in soil as discrete microcolonies whereas filamentous organisms may extend through several microhabitats. Much of the surface area of soil particles is not inhabited since the microorganisms tend to be concentrated in organic matter‐rich areas. An efficient microbial community assemblage which is commonly exploited for waste water purification (trickling filters) and fluidized beds is microbial biofilms. Biofilms are communities of microorganisms that develop at the interface of soil particles and water. They are primarily water (70 to 95 % wet weight) (Flemming, 1993) with the cells attached to the particle surface through extracellular polymers (70 to 95 % of the dry weight of the biofilm) (Flemming, 1993). Thus, the microbial cells are immobilized in a tightly organized structure, forming highly organized consortia. This physical arrangement is favorable for decomposition of complex mixtures of organic substances in systems with adequate or high nutrient levels and for optimal interaction of the community members (Costerton et al., 1995).

An additional consideration relates to the potential for a parallel conceptualization of the relationship of microorganisms in soils and the biofilms of trickling figures. That is, does a parallel community structure to the biofilm exemplified by trickling filters exist in soil and is it important for soil biological remediation. It is easy to envision the physical association of sessile microbes in soil within an assemblage akin to biofilms. The primary difficulty arises in considering the extent of the surface area of the soil particles occupied by such films. Soils tend to be carbon limited ecosystems that generally experience extremes in moisture, ranging from water saturated to highly xeric conditions. With a waste purification system, such as a trickling filter or a fluidized bed, optimization requires continuously flowing water (aerobic systems) or saturated water for those engineered processes in which anaerobic metabolism may be encouraged. These conditions (i.e., flowing oxygenated or deoxygenated water) generally do not exist in soil.

In spite of this observation, the biofilm concept can be invoked in part to explain the interaction of consortia in soil as well as the efficiency of water purification in systems such as wet lands (e.g., Alvord and Kadlec, 1996). Occurrence of highly developed biofilms is limited in most soil systems, at least for extended periods, due to the impact of biofilm development of saturated hydraulic conductivity. As the input of nutrients stimulates the development of the biofilm, the pore diameter is diminished sufficiently that saturated hydraulic conductivity can be rapidly reduced as much as four orders of magnitude (e.g., Vandevivere and Baveye, 1992), eventually reaching the point where flow of the nutrients into soil and maintenance of the biofilm community is precluded (Jennings et al., 1995). The rapidity of this biofilm development results from the tendency for colonization of the surface (inlet) of the soil column receiving the nutrient stream (see Vandevivere and Baveye, 1992). In highly structured soils, especially silt loams and clay soils, development of these pores results in the plugging of micropores. Thus, in high nutrient and water flow systems, a sandy matrix is more desirable to maintain the hydraulic conductivity (see Vandevivere et al., 1995).

Therefore, it can be concluded that the biofilm concept is useful in describing the micro‐assemblages of communities of microorganisms on surfaces of soil particles as the result of locally elevated levels of nutrients. That is, a complex, interactive microbial community certainly exists in soil microsites. A more extensive macro‐film, per se, could only be anticipated to occur under rare circumstances. The development of more traditionally described biofilm is limited to high nutrient situations, such as sites receiving chronic inputs of waste‐laden waters, a situation less likely to exist with most bioremediation situations. For bioremediation, the microbes are encountering a varied distribution of the pollutant, some free in interstitial water (as would parallel the waste stream situation to some degree), some bound or sequestered within soil micropores. Thus, both the quantity and availability of the carbon and energy source would limit the physical size of the developing biofilm.

16.2.2.3 Importance of Substrate Concentration to Remediation Success

To this point, the key elements for microbial decomposition of organic pollutants considered include the following:

· The necessity of the microbe to posses the requisite genetic potential

· The ability to express that metabolic capability in a soil ecosystem

· The requirement for the microbe to possess the appropriate traits for successful interaction with other members of the soil biological community—including the ability to become at least temporarily established therein should the use of non‐native organisms be required for successful site bioremediation

· The existence of consortia of microbial strains and species in assemblages or associations (biofilms), which optimize the capacity to decompose the complex mixtures of organic substances encountered in polluted soils which no single species or strain capable of removing the pollutant exists

A further biological trait to be addressed relates to the impact of the concentration of organic pollutant on the rate of microbial and enzymatic activities. For example, is there a concentration of the toxicant that inhibits or precludes the growth of the microbes or inhibits expression of the requisite enzymatic activities? There is an optimal range of pollutant concentrations that facilitates biological remediation. Thus, there is a need to assess our capability of managing the concentrations of these substances such that the quantities of pollutant are above the minimum required for induction and function of the enzymes required for their metabolism and below any threshold for toxicity to the decomposers. For example, the loading of the pollutant concentrations in the soil may require specific management. For example, two difficulties can be conceived to exist with excessive soil loading of toxic organic substances: a) The quantities of a carbon rich substance are sufficiently high that nitrogen or other nutrient limitation is imposed on the microbial community and/or b) The pollutant is present at levels that are toxic to the decomposer community. In either situation, dilution or reduction of the toxicant concentration is the most easily conceived, but perhaps from an engineering viewpoint, less easily achieved. An example of toxicant loading reduction followed by biodegradation is exemplified by a feasibility study of remediation of hydrocarbon contaminated soil from beneath a paint factory (Origgi et al, 1997). In that situation, bioventing and biofiltration of the outlet gas stream resulted in significant remediation of the limited area contaminated with naptha, toluene and xylenes. The residual hydrocarbons were then reduced significantly through biodegradation.

A more challenging situation relates to systems containing extremely low levels of toxicants. An environmental or ecosystem risk may exist, but the concentration of the substance is below that generally required for active microbial mineralization to occur. In this situation, microbes capable of decomposing low concentrations of the pollutant must be enriched or the quantity of the substance encountered by the microbes raised to allow for biodegradation. The latter option emerges when high levels of pollutant exist in the system but the substance has low water solubility. Thus, from the microbes’ view, only low levels of the essentially water insoluble pollutant are available for metabolism.

16.2.2.4 Managing Pollutant Water Solubility

With substances of limited water solubility, surface area is a primary determinant in controlling microbial transformations. A prime example of this situation is exemplified by the oxidation rate of elemental sulfur particles (see Chapter 14). Physical contact of the microbes with the sulfur is possible because the sulfur is not toxic. Thus, the microbes literally adhere to the sulfur particle surface. This situation is more complex when considering the decomposition of hydrophobic organic compounds. Growth on the surface of the crystalline organic may be precluded by its toxicity. Microbes existing at the surface of the crystal may actually be inhibited or killed by the capability of the hydrophobic organic contaminant to become associated with the hydrophobic portions of the cell membrane. One means of overcoming this solubility difficulty associated with hydrophobic organic substances is through the natural production of surfactants by the microbial community or use of soil amendment with industrially produced surfactants as an aspect of bioremediation plans.

Pertinent questions regarding the applicability of surfactants for the enhancement of bioavailability of soil organic contaminants of low water solubility include:

· Does the amendment of a contaminated soil with surfactants enhance the biodegradation rate of the organic pollutant?

· How does the degradation rate vary with surfactant type and concentration? That is, is more better? Are there advantages or disadvantages of anionic or nonionic surfactants?

· Are additional environmental problems created by the use of surfactants?

· If surfactant producing microbes are amended to soil, can they synthesize the surfactant and thereby enhance biodegradation of the pollutant of concern?

A variety of laboratory studies have been conducted with objectives of answering one or more of these questions. At this time, the general practicality of such research is still in question. A selection of representative studies will be examined to provide examples of the state of the art of such studies.

Evaluation of the efficacy of surfactants in the environment has been conducted using a variety of low water solubility substances, such as polycyclic aromatic hydrocarbons (PAHs), naphthalene, and xylenes, added to sediment and soil samples (e.g., Aronstein and Alexander, 1993; Grimberg et al., 1996; Liu et al., 1995; Zhang et al., 1997). In each situation, it was generally concluded that the amendment of the hydrocarbon contaminated solid matrix with surfactants stimulated the biodegradation rate of the water insoluble contaminant. Some effect of surfactant usage on both organic and inorganic soil contaminants has been noted to occur. Huang et al. (1997) found that addition of an anionic surfactant to a soil containing both lead and naphthalene resulted in enhanced solubilization of the naphthalene and desorption of the lead ion, suggesting that the surfactant could be used to increase mineralization of the hydrocarbon and to enhance the capacity to leach the lead from the soil profile.

Surfactant concentration is a significant concern for reclamation enhancement in that as its concentration increases so does the tendency to form micelles. The impact of micelle formation is that the hydrophobic pollutant may become stabilized in the micelle and thus remain unavailable to the microbial community. This was shown by Volkering et al. (1995) in their evaluation of the utility of several nonionic surfactants for enhancing the solubility of phenanthrene in batch growth experiments. In contrast, Liu et al. (1995) found that naphthalene solubilized by micelles of two nonionic surfactants was bioavailable to a mixed culture of bacteria. Zhang et al. (1997) found that phenanthrene within micells of the nonionic surfactant monorhamnolipid was less bioavailable than that contained within dirhamnolipid micelles, suggesting variation in the bioavailability due to the nature of the surfactant forming the micelles. Vanhoof and Jafvert (1996) noted that dechlorination of hexachlorobenzene varied with relationship of the surfactant level to the critical micelle concentration. For example, Tween 80 decreased dechlorination at concentrations significantly above the critical micelle concentration, whereas stimulation was noted at or below the critical micelle concentration.

Further concerns with using high concentrations of surfactant relate to the potential for toxicity of the surfactant and impact on the biological system due to the degradation of the surfactant itself. Toxicity of the surfactant is readily determined through evaluation of the effect of amendment of soil with varying concentrations of surfactant on key indicators of microbial activity, such as respiration or even mineralization of the toxicant of interest. The effect of biodegradability of the surfactant is perhaps more interesting. A study demonstrating both toxicity of surfactants and the effect of biodegradation was presented by Tiehm (1994) where it was noted that toxicity of the surfactants examined decreased with increasing hydrophobicity. The non‐toxic surfactant were effective in enhancing the degradation of a variety of hydrocarbons, fluorene, phenanthrene, anthracene, fluoranthene, and pyrene. Interestingly, sodium dodecyl sulfate inhibited the degradation of the polycyclic aromatic hydrocarbons because it was a preferred substrate. In a related study (Tiehm et al., 1997), biodegradable surfactants were found to interfere with surfactant enhancement of biodegradation of polycyclic aromatic hydrocarbons in manufactured gas plant soil due to depletion of molecular oxygen during mineralization of the surfactant.

Surfactants can be synthesized by soil microbes. Thus, it could be proposed that it could be advantageous to amend soil, not with the surfactant itself, but rather add a microbe that could survive, grow, and produce the surfactant in situ. An example of this type of system was described by Barriault and Sylvestre (1993). They amended soil contaminated with Aroclor 1242 (PCBs) with a PCB degrading bacterial strain, biphenyl as a cosubstrate to stimulate PCB degradation, and a second bacterial strain, Alcaligenes faecalis strain B‐556, which produced a surfactant. Decomposition of the Aroclor was stimulated over that not receiving the surfactant producer.

Use of soil inoculation to stimulate bioavailability of water soluble toxicants is more complex (beyond the regulatory difficulties) than simple amendment with the surfactant since a) the microbial strain producing the surfactant must be capable of survival and function within the soil system, and b) the surfactant must be produced in effective concentrations at the site where it is needed. That is, the surfactant synthesizer, the target pollutant, the organisms responsible for its decomposition, and the surfactant must all co‐exist in the same microsite at appropriate levels for optimal activity. If the decomposer population is not physically associated with the pollutant and the surfactant, then the solubolized pollutant must diffuse or be leached to the location within the soil matrix where the biodegradation can occur.

16.2.2.5 Biological Remediation of Soils with Low Loading of Pollutant

The above analysis of the complexities involved with the use of surfactants is predicated on the assumption that the soil site has been contaminated with reasonably high levels of the pollutants but that the bioavailability to the decomposer community is low due to their low water solubility. Situations may also exist where the concentrations of the organic chemical in interstitial water is limited simply because the pollutant is present at low levels. For this consideration, a low level of a pollutant is considered to be one for which there is environmental concern but the organic compound is present at levels that are insufficient to serve as a carbon and energy sources for soil microbes. In this case, the primary limitation to mineralization results from the probability of encounter of the decomposable substance and members of the decomposer population or their extracellular enzymes and the capability of the requisite enzymes to catalyze the process. The latter situation results from the fact that enzyme efficiency is optimal at substrate concentrations approximating the Michaelis constant (Km) of the enzyme. These levels are generally several orders of magnitude greater than the quantities of substrate considered in this discussion.

If the metabolizable substance is present in concentrations below that which can effectively react with the enzyme, then its stability in soil will mimic that of a recalcitrant substance unless mechanisms exist that will result in elevated concentrations in microsites. One such commonly encountered mechanism in soil results from sorption on clay surfaces. Two processes are involved in this situation. The biodecomposable pollutant is concentrated on the clay surfaces so that an elevated level of the substance exists in relationship to the soil surrounding the microsite. But, this material, due to the fact that it is sorbed on the clay surface is of limited availability to the microbial community. Since sorptive processes are reversible, this sorption of the pollutant on the clay surface also results in augmented levels in the water surrounding the clay particles. Thus, the microbe occupying the microsites, should it be capable of decomposing the pollutant, will mineralize the pollutant resulting in a reduction of the soil loading of the pollutant below levels that would be generally predicted from assessment of the pollutant concentration.

A further mechanism that can result in augmented mineralization of low level organic pollutants in soil involves stimulation of growth or activity of the decomposer population through amendment of the soil with alternate growth substrates. As was indicated above, a low level contamination of soil is defined as one in which the substance is present in concentrations less than those that are necessary to sustain microbial growth. Thus, the situation where the both the population of decomposers and the substance to be decomposed are sufficiently low that their encounter is improbable. Augmentation of the microbial population due to the presence of the pollutant can not be anticipated due to the fact that it is present in concentrations below that needed to serve as carbon and energy source. An alternative in this situation is to amend the soil with a metabolizable substance that increases the general soil microbial population (e.g., Pahm and Alexander, 1993; Schowanek et al., 1997). Schowanek et al. (1997) amended soil with tetradecenyl succinic acid (TSA), a major component of a detergent builder, with and without sewage sludge. Decomposition of the low level of TSA in the absence of sludge was preceded by a 13 day lag period whereas no lag was observed when it was accompanied by sewage sludge. Similarly, Pahm and Alexander (1993) were able to stimulate the mineralization of low concentrations of p‐nitrophenol (PNP) in soil amended with PNP degrading bacteria by the addition of glucose. Presumably, in each situation, the enhancement of the energy supply to the microbial community by the glucose and sewage sludge resulted in a sufficiently augmented energy supply for the microbial community that the pollutant concentrations could be gratuitously reduced by microbial mineralization.

16.2.3 Microbial Inoculation for Bioremediation

Given the conclusion that a soil microbial community continually adapts to its changing, evolving environment, it is logical to conclude that new members of the community are continually being recruited. These alien populations may arise from propagules that enter the soil ecosystem on soil particles carried by wind or in water or on the surfaces of invading roots. Thus, we can logically conclude that adaptation of a microbial community to chemical pollution through invasion of alien organisms should, or even must, occur, especially if soil adjacent to the contaminated site contains the appropriate decomposer populations. Therefore, the primary question regarding management of bioremediation of soil systems is not whether the potential for adaptation of the native communities to the contamination is possible. The composition of the soil microbial community must be resilient to insure survival of the ecosystem as environmental conditions evolve. The perhaps more pertinent question is: can the community development be manipulated so that highly efficient decomposer populations are selected? A logical extension of this question is: Can laboratory developed microbes (including those genetically engineered) amended to a soil system become part of the functioning soil microbial community catalyzing the effective and efficient renovation of the site?

The literature is replete with reports supporting the potential use of a variety of bacteria and fungal species for bioremediation of soil (e.g., see Madsen and Kristensen, 1997; Miethling and Karlson, 1996; Sack and Fritsche, 1997; Smith et al., 1997 as recent examples of such studies). This amendment of soil with decomposer populations may also be supplemented with a selective substrate to facilitate establishment of the inoculated population (Lajoie et al., 1993; Nishiyama et al., 1993) in situations where the host receives no or minimal advantage from the mineralization process. Furthermore, Madsen and Kristensen (1997) provide an example where a surfactant is introduced with the exogenous bacterial population to stimulate decomposition of an organic contaminant. In this latter example of introduction of microbial strains or species into a soil, polycyclic aromatic hydrocarbon decomposition was stimulated by enhanced availability of the hydrocarbon due to interaction with the surfactant produced by the laboratory cultured microbial strain. Another interesting example of such research involving managing the soil biological community for optimization of soil remediation was conducted by Siciliano and Germida (1997). They suggest that both the plant and the microbial community can be managed to enhance bioremediation.

16.2.3.1 Rhizosphere Management for Biological Remediation

When considering the interaction between plants and soil microbes from the view of enhancing bioremediation potential, the unique properties of the rhizosphere must be emphasized. Recall that the rhizosphere soil is that portion of the ecosystem where biological activity is maximized (See chapter 8). Thus, it can be generally concluded that mineralization reactions are maximized therein. Linked with the above cited studies demonstrating enhanced decomposition of organic pollutants by addition of metabolizable organic matter, the conclusion that bioremediation processes should be enhanced in the rhizosphere is logical. The question thus becomes: Are there data suggesting benefits of rhizosphere interactions in bioremediation?

A variety of studies have suggested augmented decomposition of agricultural chemicals and nonagricultural chemicals in rhizosphere soils (See Anderson et al., 1993, for a review of this topic.). For example, Reilley et al. (1996) found that plants can enhance the decomposition of polycyclic aromatic hydrocarbons in the immediate environment of the root. Their study suggested that once the toxicity of petroleum components to plants has been sufficiently reduced, incorporation of plants into a landfarming scheme can augment the bioremediation effectiveness. This stimulated activity can result simply from the general elevation of the microbial populations in the rhizosphere due to the energy and nutrient supply of the root exudates (e.g., Bollag et al., 1994). Gilbert and Crowley (1997) also note that some plant‐derived terpenoids may also promote polychlorinated biphenyl degradation by an Arthrobacter strain in the rhizosphere.

Considering that there are numerous studies demonstrating that ectomycorrhizal fungi are capable of mineralizing complex organic compounds, it is also logical to conclude that mycorrhizal associations may also be beneficial for bioremediation (e.g., Gaskin and Fletcher, 1997). For this benefit to be maximized in a managed ecosystem, propagules of the mycorrhizal fungus must be present (either through plant inoculation or as indigenous soil populations) and the plant community must develop sufficiently for these symbiotic associations to form.

A complication that must be considered when managing both microbial and plant communities for bioremediation is the potential for the formation of plant toxic intermediates during decomposition of the pollutant (e.g., Hoagland et al., 1994; Pfender, 1996). Bioremediation potential of the microbe‐plant system may be lost if toxic intermediates, which are not further metabolized, accumulate in the rhizosphere.

16.2.3.2 Distribution of Alien Organisms in Soil

In the studies cited above, the primary concern of the research was whether the organisms were capable of decomposing the target organic compound within a soil matrix. That potential activity has been adequately demonstrated in test tube studies. The next question relates to the capability of applying laboratory selected microbes to soil in a manner that allows them to pass through soil pores to the site where their action is required.

In a fluidized bed reactor, it is reasonably easy to introduce an alien microbial strain and to reach the conclusion that through the mixing process of the reactor the organism will encounter the target organic compound and mineralize or detoxify it. This capability of transporting the microbe to the site of requisite action in soil is more problematic. Generally, maintenance of as much of the native soil structure during reclamation of a soil ecosystem is desirable. Also, due to the large area requiring reclamation, mixing of microbial populations throughout the contaminated profile is generally impractical. Thus, the dynamics of microbial movement within the highly structured soil environment becomes a concern for evaluating the utility of soil inoculation for bioremediation. Considerable research has been conducted regarding the transport of various microbial strains, primarily within sand columns, and a variety of mathematical relationships have been developed describing the various transport phenomena in a variety of soil systems (e.g., see Baveye et al., 1998; Devare and Alexander, 1995; Murphy and Tate, 1996; Natsch et al., 1996). The primary conclusion from these studies is that bacteria can be transported through soil macropores preferentially during saturated flow, such as after a heavy rainfall. Limitations relate to plugging of soil pores when high populations of microbes are amended to the soil and the inability of the amended microbes to enter micropores due to size constrictions. Thus, although it can be demonstrated that with proper consideration of inoculum density and flow dynamics bacteria may be added to soil and distributed therein, the bioremediation accomplished by such processes is limited by the capability of the organisms to penetrate all microsites of the soil profile containing the pollutant. The microbes can easily be envisioned to pass through and colonize macropores and some micropores, but pollutant contained within the smaller micropores would be inaccessible to the added microbial cells.

16.2.3.3 Survival of Alien Microbes in Soil

A further concern associated with utilization of laboratory selected microbial strains for bioremediation relates to their survival beyond the time necessary for renovation of the site. It is generally considered desirable for the alien population to be eliminated once the pollutant has been exhausted or diluted to an acceptable level. Generally, to enhance biological containment of genetically modified organisms within soil systems, it is desirable to limit the survival of the organisms without limiting their beneficial effects. This process can be accomplished through the use of controlled suicide systems (e.g., Jensen et al., 1993; Knudsen and Karlstroem, 1991; Molin and Kjellber, 1993;). For example, Jensen et al. (1993) describe the utilization of two element suicide system for the containment of Pseudomonas putida KT2440. Each genetic element of the suicide system operated independently to enhance the population decline once function of the microbe was no longer required. Ahrenholtz et al. (1994) describe a system where the cell population is regulated through the destruction of intracellular DNA available for horizontal gene transfer processes.

Allied with this use of genetically engineered microbes with optimized decomposition capabilities and diminished extended survival potential (post depletion of the organic pollutant) is a consideration of our capacity for detection of the microbes. Highly sensitive techniques are required to assure detection of low populations of the amended organisms. A useful, but less sensitive, procedure involves utilization of the BIOLOG method for assessing metabolic diversity (See chapter 3). Vahjen et al. (1995) evaluated the use of this procedure for detection of Corynebacterium glutamicum in inoculated with 106 colony forming units g‐1 soil. They were able to detect the presence of the inoculated organisms as long as the cell density was greater than 105 colony forming units g‐1 soil. Matheson et al. (1997) have provided an example of using DNA probes to detect the persistence of a particular strain of Berkholderia cepacia when as few as 10 colony forming units occurred in a population of 105 non‐target colony forming units. Generally, DNA based probes are preferable for assessing survival of exogenous microbial populations in soil ecosystems.

An interesting application of genetically modified microbes in soil remediation is incorporation of fluorescence genes into the bacterial genome to provide an indication of the bioavailability of the organic pollutant within the soil system and of the degradation of the pollutant (Burlage et al., 1994; Heitzer et al., 1994; Rice et al., 1995; Webb et al., 1997). For these studies, reporter strains of bacteria are developed through incorporation of the bioluminescent lux genes from Vibrio fischeri into the genome of the genetically engineered microbe. The genes are transcriptionally fused to the catabolic gene sequence so that they are expressed whenever the catabolic genes are expressed. That is, exposure to the contaminant results in inducible bioluminescence. The magnitude of the response and the response times were shown to be concentration dependent in a system developed with Pseudomonas fluorescens for the decomposition of naphthalene and salicylate (Heitzer et al., 1994). Thus, these organisms serve as in situ indicators of the bioavailability of the pollutant (if it is present but not bioavailable, the bioluminescence genes are not induced), decomposition of the pollutant (the catabolic genes are synthesized concurrent with the bioluminescence genes), and the termination of decomposition (the genes are no longer expressed).

16.3 Soil Properties Controlling Bioremediation

The impacts of the soil physical environment on bioremediation outcome can be grouped into two categories; factors that delimit biological activity and properties that limit accessibility of the decomposer populations to the pollutant (i.e., sequestration of chemical pollutants). General analysis of the physical and chemical requirements of the soil microbial community was presented in Chapter 5. Thus, the discussion here will be limited to specific examples of ecosystem management to optimize a particular biological transformation or process.

16.3.1 Physical and Chemical Delimiters of Biological Activities

The primary soil properties managed for bioremediation considerations are soil pH (generally for the control of metal mobility), nutrient supplies (macro‐ and micronutrients as well as carbon and energy sources), and soil moisture/aeration. These delimiters of expression of biological activity provide key examples of how the rate, nature of the process, and ultimate fate of the pollutant can be selected by managing the soil physical environment. Due to the resilience of the soil microbes, fully functional microbial communities are found in what could superficially be considered to be impossible conditions. Therefore, an initial planning question for biological remediation becomes: Are the soil conditions conducive for development or existence of a fully functional microbial community? If this question is answered in the affirmative, then concerns relate to specific effects of the physical and chemical environment on the decomposition of the substance of concern (especially the biochemical processes involved) and the bioavailability of the pollutant. If not, the prevailing conditions which preclude life must be modulated prior to implementation of a management plan for control of the microbial community to optimize bioremediation.

16.3.1.1 pH

In the previous consideration of the effect of soil pH on microbial activity (Chapter 5), it was noted that microbial growth and activity occurs in both acidic and alkaline soil ecosystems. Therefore, there is no inherently bad implication of the occurrence of soil systems at the ranges of pH values commonly beyond the ranges normally encountered for agricultural production in temperate regions (slightly acidic to neutral pH). In fact, numerous tropical cropping systems (pineapple, cassava, etc.) are productive on soils with pH’s of less than 4. The major requirement for reclamation management is that the full implications of pH on the processes occurring in the chemically polluted soil be considered. Examples of soil processes where pH effects need to be evaluated are nutrient cycling and effects of pH on the physical and chemical states of the organic chemical of question.

Nutrient cycling processes most commonly managed for bioremediation relate to nitrogen availability. Note that it was indicated in chapter 11 that both nitrification and symbiotic nitrogen fixation can be impacted by acidic conditions. Autotrophic nitrifiers function above pH 5.0 to 5.5 whereas the Rhizobium‐legume symbiotic associations for nitrogen fixation require a soil pH greater than about 5.5. Thus, for example, in acidic soils with low levels of fixed nitrogen, the soils will need to be limed to a pH level conducive for nitrogen fixation to occur with legumes, or actinorhizal associations, such as bayberry, have to be cultivated.

The greater concern involving soil pH is likely the impact of pH on the ionic state of the pollutant and the resultant effect of this charge variation on its interaction with the enzymes involved in its mineralization or detoxification as well as its sorption to soil organic matter and clays. At pH values above the pKa, acidic groups are ionized, facilitating ionic interactions with enzymes as well as the physical environment whereas below the pKa, hydrophobic interactions are favored. Similarly, from an environmental health view, hydrophobic interactions tend to favor associations with membrane lipids and bioaccumulation of the toxicants. An example of the effect of pH variation on the state of an organic chemical in soil is provided by looking at the sorptive properties of atrazine and fomesafen, two pesticides, in soil (Weber, 1993). Decreasing pH resulted in higher sorption of both herbicides. Atrazine sorption resulted from ionic bonding at low pH and physical bonding at neutral pH whereas fomesafen sorption was the product of physical forces at neutral pH and hydrophobic bonding and/or precipitation at low pH. Thus, at lower pHs, the bioavailability of the pesticides could be anticipated to be less than at high or neutral pHs.

16.3.1.2 Nutrient Amendment and Bioremediation

Examples of nutrient amendments for bioremediation of soils are generally associated with hydrocarbon rich contaminants, such as petroleum. It has long been known that nitrogen fertilization is necessary to stimulate petroleum degradative with land farming. Even in situations where other variables such as temperature in alpine soils (Margesin and Schinner, 1997) or salinity and temperature with beaches of Prince William Sound (Pritchard, 1991), use of nitrogenous fertilizers enhanced biodegradation. Other nutrient amendments may also involve provision of a carbon and energy sources in situations where the pollutant can only be decomposed cometabolically (e.g., carbon tetrachloride and a variety of easily metabolized carbon sources (Lewis and Crawford, 1993)) or simply from the general stimulation of the microbial population (as indicated above). In the latter situation, composted waste materials provide an energy source for the general soil microbial populations (thereby stimulating mineralization of the contaminant), stimulate soil aggregate forming processes resulting in an improved environment for function of the microbial community, improve the soil moisture dynamics of the system, and also result in increased humification of the organic toxicant, thereby providing an alternative path for removal of the toxicant from the system (e.g., Kastner et al., 1995; Laine and Jorgensen, 1996). Thus, recycling of composted societal by‐products can assist in optimization of soil bioremediation processes.

16.3.1.3 Soil Moisture/Aeration

As was indicated in Chapter 5, the world of soil microbes can be conceptualized as an aquatic system, albeit a thin film of water covering a largely particulate matrix. The organisms are bathed in this film of water, nutrients are transported into and away from the cell by diffusion or mass transport processes, and the water itself is a requisite participant in many biochemical transformations. These observations are true whether we are dealing with a native soil ecosystem or a remediation site. Of particulate concern in managing polluted soil is the interrelationship between soil moisture and soil oxygen status. That is, decomposition rates, detoxification processes, the extent of mineralization, and the nature of the metabolic intermediates accumulating vary with the nature of the terminal oxygen acceptor. Much bioremediation philosophy has been predicated on the objective of optimizing free oxygen levels. This management strategy is based on the assumption that oxygen based metabolism is more rapid and generally leads to accumulation of fewer metabolic intermediates than would occur under anoxic conditions. Considering that the terrestrial system is bathed in an oxygen rich atmosphere, this philosophy is generally reasonable, although as will be indicated below, consideration of other terminal electron acceptors (e.g., nitrate, sulfate, carbon dioxide) is useful in naturally anoxic systems where manipulation of aerobic processes is impractical.

The examination of native soil ecosystems, such as swamps and related water saturated systems, leads to a conclusion that organic matter decomposition is slower and the accumulation of partially decomposed organic intermediates is accentuated in such systems. It has long been known that the versatility of the soil microbial community in dealing with reduced oxygen levels is revealed in such systems. For example, reduction of oxygen tensions a percent or two from the normal atmospheric level of approximately 21.0 % is limiting for higher animals, but in drained histosols, microbial mineralization of the accumulated organic matter continues to free oxygen tensions of less than 1 % (see Broadbent, 1960). Similar relationships of biodegradation processes to available oxygen levels would be anticipated to occur in soils contaminated with organic chemicals, although the rate of decomposition will necessarily vary directly with oxygen tension as the oxygen levels become limiting. Examples of limitations to bioremediation by oxygen levels are common in studies of soil and sediment remediation processes (e.g., Hurst et al., 1997; Madsen et al., 1996). Hurst et al. (1997) noted that pentachlorophenol degradation in contaminated soil from a superfund site was precluded in the absence of oxygen, but they were enhanced by soil oxygen concentrations between 2 and 21%. They concluded that a soil oxygen concentration of at least 2 to 5% was necessary for pentachlorophenol bioremediation. When depletion of soil oxygen levels by biological activity exceeds the diffusion rate for oxygen replenishment, the oxygen tension can be elevated through bioventing (e.g., Hurst et al., 1997) or even amendment of the soil with peroxides (e.g., Fiorenza and Ward, 1997).

In some situations, soil oxygen can be depleted to optimize processes favored by anoxic conditions (e.g., dehalogenation or reduction of nitro groups). Kaake et al. (1992) describe a soil treatment method for removal of the pesticide dinoseb (2‐sec‐butyl‐4,6‐dinitrophenol) where soils were pretreated by amendment with a starchy potato processing by‐product to reduce the redox potential and create anaerobic conditions. An anaerobic microbial consortium capable of mineralizing dinoseb without formation of polymerization products commonly seen under aerobic or microaerobic conditions developed. Alternatively, alternation of aerobic and anoxic conditions can be used to facilitate bioremediation. For example, a two stage process starting with an anaerobic conditions facilitates the bioremediation of munitions compounds including 2,4,5‐trinitrotoluene (TNT) (e.g., Funk et al. 1993; Breitung et al. 1996). During the first phase, ammunition residues and aminoaromatic compounds were depleted and the subsequent reduced products were mineralized during the aerobic phase of the process.

It is a given that the bulk of organic matter decomposition, by far, in soil results from oxygen based processes. Thus, it is logical that initial considerations in developing bioremediation plans must involve evaluation of the potential for optimization of aerobic soil metabolism; but, there are many contaminated soil ecosystems where utilization of alternate electron acceptors is more practical and economical due to the fact that the site is naturally anoxic. A further encouragement in this consideration is the fact that in many instances the soil contaminants may include nitrate or sulfate, two logical terminal electron acceptor alternatives. Furthermore, dissimilatory nitrate and sulfate reduction yield energy recoveries to the microbes approximating those gained from oxygen metabolism. Thus, it is reasonable to conclude that these processes should be as efficient and rapid as the more commonly studied and applied aerobic degradation processes.

Denitrification is the most studied anoxic process for bioremediation. Denitrification, as a means of reducing soil nitrate loading, by definition is a bioremediation process (e.g., see Smith et al., 1994; Weier et al., 1994). Since a carbonaceous compound is generally mineralized as the source of electrons for denitrification, a potential exists for selection of denitrifiers that mineralize toxic organic compounds for bioremediation of soils contaminated with both nitrate and decomposable organic chemicals (see Casella and Payne (1996) for a review of this topic.) Indeed, as indicated above, a variety of potential soil toxicants are mineralized by denitrifiers (see also Crawford et al., 1998; Haner et al., 1995). A major question remaining relates to the capability of managing bioremediation in field situations under denitrifying conditions. An example of the potential for manipulation of anoxic processes is provided by bioremediation of BTEX (benzene, toluene, ethylbenzene, and xylene) in petroleum contaminated soils. BTEX degradation has been noted with ferric iron, sulfate or nitrate as terminal electron acceptors as well as under methanogenic conditions (Lovley, 1997). Denitrification alone is apparently not sufficient to bioremediate gasoline (BTEX) contaminated soils due to the inability of the denitrifiers to decompose benzene (Kao and Borden, 1997).

16.3.2 Sequestration and Sorption Limitations to Bioavailability

A real limitation to the successful conclusion of biological remediation efforts relates to the quantity of residual pollutant remaining in the system. The longer the time interval since initial contamination of the soil, the greater the probability that significant quantities of the organic pollutant will not be accessible to the biological community. A key to understanding this limitation to the ultimate cleanup of the site resides in developing an understanding of the impact of soil heterogeneity on the biological processes occurring therein. As was emphasized in Chapter 1, the variety of minerals, complex organic substances, as well as biological entities existent in any soil results in a myriad of different microenvironments within which soil microbes function. This heterogeneity affects both the immediate and potential bioavailability of decomposable organic substances and toxicants. For an organic compound to be decomposed or for this substance to inhibit microbial activity, it must have contact with the microbe or enzymes produced by the microbe. (Recall that extracellular enzymes may transform organic matter into water soluble products, which through mass transport or diffusion come into direct contact with the microbial cell.) Two primary properties of the soil system that affect bioavailability are physical location within the soil pore structure and sorption onto soil particulate surfaces.

Due to the dynamic nature of the distribution of materials within the soil matrix, the quantity of substance sequestered within soil pores, sorbed to soil minerals or dissolved in soil water changes with time. It is commonly observed that the bioavailability, chemical extractability, and decomposition rate of soil organic contaminants decreases with time since initial exposure (e.g., Alexander, 1995; Grant et al., 1995; Radosevich, 1997). This time dependent variation in bioavailability can be demonstrated through assessment of percent uptake by earthworms, bacterial degradation, or solvent extractability (e.g, Kelsey et al., 1997). For example, Kelsey and Alexander (1997) found that less rigorous solvent extraction methods correlated better with the availability of atrazine, phenanthrene and napthalene to earthworms in sterile soil than did the more vigorous extraction procedures commonly used to assess soil contamination. The degree of sequestration varies with time as well as with soil properties, such as clay and organic matter contents (e.g., White et al., 1997). Furthermore, the source of the microbial inoculum may also affect the extent of biodegradation that occurs in the soil system. Sandoli et al. (1996) found that microbes isolated from their sand sediments that were contaminated with phenthrene decomposed the contaminant with a shorter lag period than did those from other sources.

These observations support the hypothesis that the bioavailability of organic substances in soil to decomposition by soil microbes relates to the pore structure of the soil as well as the capacity of the substances to interact with clay minerals and colloidal organic matter. This hypothesis has been supported by use of model systems composed of glass, polystyrene, diatomite, and silica beads with varying pore structure and hydrophobicity (Nam and Alexander, 1998). This study demonstrated that both hydrophobicity and existence of nanopores in the solid matrix affected the bioavailability of phenanthrene.

It is reasonably easy to envision the nature of the effect of micropore structure on chemical‐microbe interactions. Substances trapped within soil pores of mean diameters less than that of microbes and containing no microbes would not be detoxified or mineralized. The quantity of pollutant retained in micropores is controlled by the physical forces dictating soil water dynamics. Entrance into the pores is diffusion limited, as opposed to the soil macropores where saturated flow allows for mass transport of solutes in gravitational water. Once solutes enter the micropores their exit from the system is controlled by a) diffusion, b) their retention within the more rigid crystalline water layer on the pore surface, and c) their tendency to sorb onto the surfaces of clay or organic matter surfaces in the pores. The role of sorptive phenomena is documented by the observation that biodegradation in soil containing both pollutant and its decomposers is frequently proportional to the desorption rate of the pollutant (Carmichael et al., 1997). The fact that the nature of the mineral component of the soil (especially clay type) affects decomposition kinetics underscores the need to evaluate the soil mineralogy (e.g., Appitz and Meyersschulte, 1996; Handerlein et al., 1996). Non‐covalent interactions of organic substances with soil humic substances can result in associations with stability resembling that of covalently bound substances (Dec and Bollag, 1997). Chemical models have shown that simple organic compounds can form stable associations with clay surfaces and within pores of soil humic acids that are sufficiently stable to essentially remove the toxicant or biodegradable substance for the biologically available nutrient pool (Schulten, 1995; Schulten and Schnitzer, 1997). Thus, bioavailability of organic chemicals in soil and their potential for biological remediation is controlled by the tendency of the substance to diffuse into and out of soil micropores as well as by the forces preventing diffusion of the chemical back out of the pore (sorption, non‐covalent stabilization in humic acids and retention in crystalline water.)

16.4 Concluding Observations

Bioremediation of polluted soils is only a possibility because of the extensive diversity of the soil microbial community and the redundancy of capabilities that exists therein. Indeed, even though there is a potential to develop laboratory derived microbial species and strains capable of decomposing soil pollutants in situ, the bulk of soil biological remediation is facilitated or accomplished by the indigenous soil microbial community. Thus, success of any remediation project relies totally on a clear understanding of the biological capability inherent in the system and the physical, chemical, and biological limitations of the expression of that activity. A variety of soil properties can be varied to optimize the rate and extent of biological remediation, but the ultimate controller of the quantity of toxicant remaining in the system is its bioavailability. Retention of chemicals within soil micropores and association with soil minerals and organic matter precludes total elimination of the soil contaminant. The ultimate environmental risk of this residual toxicant is determined by the potentiality for the bioavailability to change with time. The primary unanswered question in relationship to the overall risk of retaining sequestered toxicants in a remediated soil relates to the still to be determined probability that the toxicant will be released in meaningful concentrations at a future date. Based on the above analysis of the limitations to bioremediation, a meaningful concentration would be one that poses a toxic threat to the function of the indigenous biological community or to higher plants or animals in the ecosystem.

References

1. Ahrenholtz, I., M. G. Lorenz, and W. Wackernage. 1994. A conditional suicide system in Escherichia coli based on the intracellular degradation of DNA. Appl. Environ. Microbiol. 60:3746–3751.

2. Alexander, M. 1995. How toxic are toxic chemicals in soil. Environ. Sci. Technol. 29:2713–2717.

3. Alvord, H. H., and R. H. Kadlec. 1996. Atrazine fate and transport in the Des Plaines wetlands. Ecol. Modelling 90:97–107.

4. Anderson, T. A., E. A. Guthrie, and B. T. Walton. 1993. Bioremediation in the rhizosphere. Environ. Sci. Technol. 27:2630–2636.

5. Apitz, S. E., and K. J. Meyersschulte. 1996. Effects of substrate mineralogy on the biodegradability of fuel components. Environ. Toxicol. Chem. 15:1883–1893.

6. Aronstein, B. N, and M. Alexander. 1993. Effect of a non‐ionic surfactant added to the soil surface on the biodegradation of aromatic hydrocarbons within the soil. Appl. Environ. Microbiol. 39:386–390.

7. Barriault, D., and M. Sylvestre. 1993. Factors affecting PCB degradation by an implanted bacterial strain in soil microcosms. Can. J. Microbiol. 39:594–602.

8. Baveye, P., P. Vandevivere, B. L. Hoyle, P. C. DeLeo and D. Sanchez de Lozda. 1998. Environmental impact and mechanism of biological clogging of saturated soils and aquifer materials. Crit. Rev. Environ. Sci. Technol. In Press.

9. Bedard. D. L., and J. F. Quensen III. 1995. Microbial reductive dechlorination of polychlorinated biphenyls. In L. Y. Young and C. E. Cerniglia (eds.) Microbial transformation and Degradation of Toxic Organic Chemicals. P. 127–216. Wiley‐Liss, NY.

10. Bollag, J.‐M., T. Mertz, and L. Otjen. 1994. Role of microorganisms in soil bioremediation. Bioremediation through rhizosphere Technology, American chemical Society Symposium Series 563:2–10.

11. Broadbent, F. E. 1960. Factors influencing the decomposition of organic soils of the California Delta. Hilgardia 29:587–612.

12. Breitung, J., D. Bruns‐Nagel, K. Steinbach, L. Kaminski, D. Gemsa, and E. von Low. 1996. Bioremediation of 2,4,6‐trinitrotoluene‐contaminated soils by two different aerated compost systems. Appl. Microbiol. Biotechnol. 44:795–800.

13. Burke, D. J., K. A. Smemo, J. C. López‐Gutiérrez and J. L. DeForest. 2012. Soil fungi influence the distribution of microbial functional groups that mediate forest greenhouse gase emissions. Soil Biol. Biochem. 53:112–119.

14. Burlage, R. S., A. V. Palumbo, A. Heitzer, and G. Sayler. 1994. Bioluminescent reporter bacteria detect contaminants in soil samples. Appl. Biochem. Biotechnol. 45:713–740.

15. Casella, S., and W. J. Payne. 1996. Potential of denitrifiers for soil environment protection. FEMS Microbiol. Lett. 140:1–8.

16. Carmichael, L. M., R. F. Christman, and F. K. Pfaender. 1997. Desorption and mineralization kinetics of phenanthrene and chrysene in contaminated soils. Environ. Sci. Technol. 31:126–132.

17. Cartmill, A. D., D. L. Cartmill. 2014. Short‐term biodegradation of Petroleum in Planted and Unplanted Sandy Soil. J. Environ. Qual. 42: 1080–1085.

18. Chen, T. C. Phillips, J. Hamilton, B. Chartbrand, J. Grosskig, K. Bradshaw, T. Carlson, K. Timlick, D. Peak, and S. D. Siciliano. 2014. Citrate addition increased phosphorus bioavailability and enhanced gasoline bioremediation. J. Eviron. Quality: 46:975–983.

19. Clothier, B. E., and S. R. Green. 1997. Roots: The big movers of water and chemical in soil. Soil Science 162:534–543.

20. Comerford, N. B., Franzluebbers, M. E. Stromberger, L. Morris, D. Markewitz, and R. Moore. (2013) Assessment and evaluation of soil ecosystem Services. Soil Horizons. 54:1–54

21. Costerton, J. W., Z. lewandowski, d. E. Caldwell, D. R. Korer, and H. M. Lappin‐Scott. 1995. Microbial biofilm. Annu. Rev. Microbiol. 49:711–745.

22. Crawford, J. J., G. K. Sims, r. L. Mulvaney, and M. Radosevich. 1998. Biodegradation of atrazine under denitrifying conditions. Appl. Microbiol. Biotechnol. 49:618–623.

23. Devare, M., and M. Alexander. 1995. Bacterial transport and phenanthrene biodegradation in soil and aquifer sand. Soil Sci. Soc. Am. J. 49:1316–1320.

24. Dec, J., and J.‐M. Bollag. 1997. Determination of colvalent and noncovalent binding interactions between xenobiotic chemicals and soil. Soil Sci. 162:858–874.

25. Field, J. A., J. M. Stams, M. Kato, and G. Shraa. 1995. Enhanced biodegradation of aromatic pollutants in cocultures of anaerobic and aerobic bacterial consortia. Antonie van Leeuwenhoek. 67:47–77.

26. Fiorenza. S., and C. H. Ward. 1997. Microbial adaptation to hydrogen peroxide and biodegradation of hydrocarbons. J. Indust. Microbiol. Biotechnol. 18:140–151.

27. Flemming, H. C. 1993. Biofilms and environmental protection. Water Sci. Technol. 27:1–10.

28. Fliermans, C. B., T. J. Phelps, D. Ringelberg, A. T. Mikell, and D. C. White. 1988. Mineralization of trichloroethylene by heterotrophic enrichment cultures. Appl. Environ. Microbiol. 54:1709–1714.

29. Funk, S. B., D. J. Roberts, D. L. Crawford, and R. L. Crawford. 1993. Initial‐phase optimization for bioremediation of munition compound‐contaminated soils. Appl. Environ. Microbiol. 59:2171–2177.

30. Gaskin, J. L., and J. fletcher. 1997. The metabolism of exogenously provided atrazine by the ectomycorrhizal fungus Hebeloma crustuliniforme and the host plant Pinus ponderosa. Phytoremediation of Soil and Water Contaminants, American Chemical Society Symposium Series. 664:152–160.

31. Gilbert, E. S., and D. E. Crowley. 1997. Plant compounds that induce polychlorinated biphenyl biodegradation by Arthrobacter sp. Strain B1B. Appl. Environ. Microbiol. 63:1933–1938.

32. Grant, C. L., T. F. Jemloms, K. F. Meyers, and E. F. McCormick. 1995. Holding‐time estimates for soils containing explosives residues – comparisons of fortification vs field contamination. Environ. Toxicol. Chem. 14:1865–1874.

33. Grimberg, S. J., W. T. Stringfellow, and M. D. Aitken. 1996. Quantifying the biodegradation of phenanthrene by Pseudomonas stutzeri P16 in the presence of a nonionic surfactant. Appl. Environ. Microbiol. 62:2387–2392.

34. Haderlein, S. B., K. W. Weissmahr, and R. P. Schwarzenbach. 1996. Specific adsorption of nitroaromatic – explosives and pesticides to clay mierals. Environ. Sci. Technol. 30:612–622.

35. Haner, A., P. Hohener, and J. Zeyer. 1995. Degradation of p‐xylene by a denitrifying enrichment culture. Appl. Environ. Microbiol. 61:3185–3188.

36. Heitzer, A., K. Malachowsky, J. E. Thonnard, P. R. Bienkowski, D. C. White, and G. S. Sayler. 1994. Optical biosensor for environmental on‐line monitoring of naphthalene and salicylate bioavailabiltiy with an immobilized bioluminescent catabolic reporter bacterium. Appl. Environ. Microbiol. 60:1487–1494.

37. Hoagland, R. E., R. M. Zablotowice, and M. E. Locke. 1994, Propanil metabolism by rhizosphere microflora. Pp. 160—183. In bioremediation through rhizosphere technology. ACS Symposium Series, American Chemical Society, Washington D. C.

38. Huang, C., J. E. Vanbenschoten, T. C. Healy, and M. E. Ryan. 1997. Feasibility study of surfactant use for remediation of organic and metal contaminated soils. J. Soil contamin. 6:537–556.

39. Huesemann, M. H. 1997. Incomplete hydrocarbon biodegradation in contaminated soils: Limitations in bioavailability of inherent recalcitrance. Bioremed. J. 1:27–39.

40. Hurst, C. J. R. C. Sims, J. L. Sims, D. L. Sorensen, J. E. Mclean, S. Huling. 1997. Soil gas oxygen tension and pentachlorophenol biodegradation. J. Environ. Engin. 123:364–370.

41. Jennings, D. A., J. N. Petersen, R. S. Skeen, B. S. Hooker, B. M. Peyton, D. L. Johnstone, and D. R.Yonge. 1995. Effects of variations in nutrient loadings on pore plugging in soil columns. Appl. Biochem. Biotech. 51/52:727–734.

42. Jensen, L. B., J. L. Ramos, Z. Kaneva, and S. Molin. 1993, A substrate‐dependent biological containment system for Pseudomonas putida based on the Escherichia coli gef gene. Appl. Environ. Microbiol. 59:3713–3717.

43. Juteau, P., R. Beaudet, G. McSween, F. Lepine, S. Milot, and J.‐G. Bisaillon. 1995. Anaerobic biodegradation of pentachlorophenol by a methanogenic consortium. Appl. Microbiol. Biotech. 44:218–224.

44. Kaake, R. H., D. G. Roberts, T. O. Stevens, T. O. Crawford, and D. L. Crawford. 1992. Bioremediation of soils contaminated with the herbicide 2‐sec‐butyl‐4,6‐dintirophenol (dinoseb). Appl. Environ. Microbiol. 58:1683–1689.

45. Kao, C. M. and R. C. Borden. 1997. Site‐specific variability in BTEX biodegradation under denitrifying conditions. Groundwater 35:305–311.

46. Kastner, M., S. Lotter, J. Heerenklage, M. Breuer‐Jammali, R. Stegmann, and B. Mahro. 1995. Fate of 14C‐labeled anthracene and hexadecane in compost manured soil. Appl. Microbiol. Biotechnol. 43:1128–1135.

47. Kazumi, J., M. M. Häggblom, and L. Y. Young. 1995. Diversity of anaerobic microbial processes in chlorobenzoate degradation: Nitrate, iron, sulfate and carbonate as electron acceptors. Appl. Microbiol. Biotech. 43:929–936.

48. Kelsey, J. W., and M. Alexander. 1997. Declining bioavailability and inappropriate estimation of risk of persistent compounds. Environ. Toxicol. Chem. 16:582–585.

49. Kelsey, J. W., B. D. Kottler, and M. Alexander. 1997. Selective chemical extractants to predict bioavailability of soil‐aged organic chemicals. Environ. Sci. Technol. 31:214–217.

50. Knudsen, S. M, and O. H. Karlstoem. 1991. Development of efficient suicide mechanisms for biological containment of bacteria. Appl. Environ. Microbiol. 57:85–92.

51. Laine, M. M., and K. S. Jorgensen. 1996. Straw compost and bioremediated soil as inocula for the bioremediation of chlorophenol‐contaminated soil. Appl. Environ. Microbiol. 62:1507–1513.

52. Lajoie, C. A., G. J. Zylstra, M. F. DeFlaun, and P. F. Strom. 1993. Development of field application vectors for bioremediation of soils contaminated with polychlorinated biphenyls. Appl. Environ. Microbiol. 59:1735–1741.

53. Lewis, T. A., and R. L. Crawford. 1993. Physiological factors affecting carbon tetrachloride dehalogenation by the denitrifying bacterium Pseudomonas sp. Strain KC. Appl. Environ. Microbiol. 59:1635–1641.

54. Liu, Z. B., A. M. Jacobson, and R. G. Luthy. 1995. Biodegradation of naphthalene in aqueous nonionic surfactant systems. Appl. Environ. Microbiol. 61:145–151.

55. Lovley, D. R. 1997. Potential for anaerobic bioremediation of BTEX in petroleum‐contaminated aquifers. J. Indust. Microbiol. Biotechnol. 18:75–81.

56. Madsen, T., and P. Kristensen. 1997. Effects of bacterial inoculation and nonionic surfactants on degradation of polycyclic aromatic hydrocarbons in soil. Environ. Toxicol. Chem. 16:631–637.

57. Madsen, E. L., C. L. Mann, and S. E. Bilotta. 1996. Oxygen limitations and aging as explanations for the field persistence of naphthalene in coal tar‐contaminated surface sediments. Environ. Toxicol. Chem. 15:1876–1882.

58. Margesin, R., and F. Schinner. 1997. Efficiency of indigenous and inoculated cold‐adapted soil microorganisms for biodegradation of diesel oil in alpine soils. Appl. Environ. Microbiol. 63:2660–2664.

59. Matheson, V. G., J. Manakata‐Marr, G. D. Hopkins, P. L. McCarthy, J. M. Tiedje, and L. J. Forney. 1997. A novel means to develop strains‐specific DNA probes for detecting bacteria in the environment. Appl. Environ. Microbiol. 63:2863–2869.

60. Miethling, R., and U. Karlson. 1996. Accelerated mineralization of pentachlorophenol in soil upon inoculation with Mycobacterium chlorophenolicum PCP1 and Sphingomonas chlorophenolica RA2. Appl. Environ. Microbiol. 62:4361–4366.

61. Molin, S., and S. Kjellberg. 1993. Release of engineered microorganisms: Biological containment and improved predictability for risk assessment. Ambio 22:242–245.

62. Murphy, S. L., and R. L. Tate III. 1996. Bacterial movement through soil. in G. Stotzky and J. M. Bollag (eds.) Soil Biochem. 9:253–286. Dekker, NY.

63. Nam, K., and M. Alexander. 1998. Role of nanoporosity and hydrophobicity in sequestration and bioavailability: Tests with model solids. Environ. Sci. Technol. 32:71–74.

64. Natsch, A., C. Keel, J. Troxler, M. Zala, N. Von Albertini, an G. Defago. 1996. Importance of preferential flow and soil management in vertical transport of a biocontrol strain of Pseudomonas fluorescens in structured field soil. Appl. Environ. Microbiol. 62:33–40.

65. Nishiyama, M., K. Senoo, and S. Matsumoto. 1993. Establishment of γ‐1,2,3,4,5,6‐hexachlorocyclohexane‐assimilating bacterium, Sphingomonas paucimobilis strain SS86 in soil. Soil Biol. Biochem. 25:769–774.

66. Origgi, G., M. Colombo, F. Depalma, M. Rivolta, P. rossi and V. Andreoni. 1997. Bioventing of hydrocarbon‐contaminated soil and biofiltration of the off‐gas‐ Results of a field scale investigation. J. Environ. Sci. & Hlth. 32:2289–2310.

67. Pahm, M. A., and M. Alexander. 1993. Selecting inocula for the biodegradation of organic compounds at low concentrations. Microbial Ecol. 25:275–286.

68. Pfender, W. F. 1996. Bioremediation bacteria to protect plants in pentachlorophenol‐contaminated soil. J. Environ. Qual. 25:1256–1260.

69. Pritchard, P.H. 1991. Bioremediationas a technology: Experiences with the Exxon Valdez oil spill. Ecological Research Series. U.S. Environmental Protection Agency. Washington, DC.

70. Radosevich, M., S. J. Traina, and O. H. Tuovinen. 1997. Atrazine mineralization in laboratory‐aged soil microcosms inoculated with S‐triazine‐degrading bacteria. J. Environ. Qual. 26:206–214.

71. Reilley, K. A., M. K. Banks, and A. P. Schwab. 1996. Dissipation of polycyclic aromatic hydrocarbons in the rhizosphere. J. Environ. Qual. 25:212–219.

72. Rice, J. F., R. F. Fowler, A. A. Arrage, D. C. White, and G. S. Sayler. 1995. Effects of external stimuli on environmental bacterial strains harboring an algD‐lux bioluminsecent reporter plasmid for the study of corrosive biofilms. J. Industr. Microbiol. 15:318–328.

73. Sack, U., and W. Fritsche. 1997. Enhancement of pyrene mineralization in soil by wood‐decaying fungi. FEMS Microbiol. Ecol. 22:77–83.

74. Sandoli, R. L., W. C. Ghiorse, and E. L. Madsen. 1996. Regulation of microbial phenanthrene mineralization in sediment samples by sorbent‐sorbate contact time, inocula and gamma irradiation‐induced sterilization artifacts. Environ. Toxicol. Chem. 15:1901–1907.

75. Schowanek, D. R., T. C. J. Feijtel, and T. W. Federle. 1997. Effect of concentration and environmental form of tetradecenyl succinic acid on its mineralization in soil. Biodegradation 7:377–382.

76. Schulten, H.‐R. 1995. The three‐dimensional structure of soil organo‐mineral complexes studyied by analytical pyrolysis. J. Anal. Appl. Pyrol 32:111–126.

77. Schulten, H.‐R., and M. Schnitzer. 1997. Chemical model structures for soil organic matter and soils. Soil sci. 162:115–130.

78. Siciliano, S. D., and J. J. Germida. 1997. Bacterial inoculants of forage grasses that enhance degradation of 2‐chlorobenzoic acid in soil Environ. Toxicol. Chem. 16:1098–1104.

79. Smith, M. J., G. Lethbridge, and R. G. Burns. 1997. Bioavailability and biodegradation of polycyclic aromatic hydrocarbons in soils. FEMS Microbiol. Letters 152:141–147.

80. Smith, R. L., M. L. Ceazan, and M. H. Brooks. 1994. Autotrophic hydrogen‐oxidizing, denitrifying bacteria in groundwater, potential agents for bioremediation of nitrate contamination. Appl. Environ. Microbiol. 60:1949–1955.

81. Stehmeier, L. G., T. R. Jack, and G. Voordouw. 1996. In vitro degradation of dicyclopentadiene by microbial consortia isolated from hydrocarbon contaminated soil. Can. J. Microbiol. 42:1051–1060.

82. Sutherland, J. B., F. Rafii, A. A. Khan, and C. E. Cerniglia. 1995. Mechanisms of polycyclic aromatic hydrocarbon degradation. In L. Y. Young and C. E. Cerniglia (eds.) Microbial Transformation and Degradation of Toxic Organic Chemicals. P. 269–306. Wiley‐Liss, NY.

83. Swobada‐Colberg, N. G. 1995. Chemical contamination of the environment: Sources, types, and fate of synthetic chemicals. In L. Y. Young and C. E. Cerniglia (eds.) Microbial Transformation and Degradation of Toxic Organic Chemicals. P. 27–74. Wiley‐Liss, NY.

84. Tiehm, A. 1994. Degradation of polycyclic aromatic hydrocarbons in the presence of synthetic surfactants. Appl. Environ. Microbiol. 60:258–263.

85. Tiehm, A., M. Stieber, P. Werner, and F. H. Frimmel. 1997. Surfactant‐enhanced mobilization and biodegradation o polycyclic aromatic hydrocarbons in manufactured gas plant soil. Environ. Sci. Techn. 31:2570–2576.

86. Vahjen, W., J.‐C. Munch, and C. C. Tebbe. 1995. Carbon source utilization of soil extracted microorganisms as a tool to detect the effects of soil supplemented with genetically engineered and non‐engineered Corynebacterium glutamicum and a recombinant peptide at the community level. FEMS Microbiol. Ecol. 18:317–328.

87. Vandevivere, P., and P. Baveye. 1992. Saturated hydraulic conductivity reduction caused by aerobic bacteria in sand columns. Soil Sci. Am. J. 56:1–13.

88. Vandevivere, P., P. Baveye, D. Sanchez de Lozada, and P. DeLeo. 1995. Microbial clogging of saturated soils and aquifer materials: Evaluation of mathematical models. Water Resources Res. 31:2173–2180.

89. Vanhoof, P. L., and C. T. Jafvert. 1996. Reductive dechlorination of chlorobenzenes in surfactant‐amended sediment slurries. Environ. Toxicol. Chem. 15:1814–1924.

90. Volkering, F., A. M. Breure, J. G. Vanandel, and W. H. Rulkens. 1995. Influence of nonionic surfactants on bioavailability and biodegradation of polycyclic aromatic hydrocarbons. Appl. Environ. Microbiol. 61:1699–1705.

91. Wackett, L. P. 1995. Bacterial co‐metabolism of halogenated organic compounds. In L. Y. Young, and C. E. Cerniglia (eds.) Microbial Transformation and Degradation of Toxic Organic Chemicals. P. 217–241. Wiley‐Liss, NY.

92. Webb, O. F., P. R. Bienkowsi, U. Matrubutham, F. A. Evans, A. Heitzer, and G. S. Sayler. 1997. Kinetics and response of a Pseudomonas fluorescens HK66 biosensor. Biotechnol. Bioeng. 54:491–502.

93. Weber, J. B. 1993. Ionization and sorption of fomesafen and atrazine by soils and soil constituents. Pesticide Sci. 39:31–38.

94. Weier, K. L., J. W. Doran, A. R. Mosier, J. F. Power, and T. A. Peterson. 1994. Potential for bioremediation of high nitrate irrigation water via denitrification. J. Environ. Qual. 23:105–110.

95. White, J. C., J. W. Kelsey, P. B. Hatzinger, and M. Alexander. 1997. Factors affecting sequestration and bioavailability of phenanthrene in soils. Environ. Toxicol. Chem. 16:2040–2045.

96. Xie, H., T. F. Guetzloff, and J. A. Rice. 1997. Fractionation of pesticide residues bound to humin. Soil Sci. 162:421–429.

97. Young, L.Y., and C. E. Cerniglia (eds.) 1995. Microbial Transformations and Degradation of Toxic Organic Chemicals. John Wiley & Sons, Inc. NY. 654 pp.

98. Zhang, Y. M., W. J. Maier., and R. M. Miller. 1997. Effect of rhamnolipids on the dissolution, bioavailabiltiy and biodegradation of phenanthrene. Environ. Sci. Technol. 31:2211–2217.

99. Zouari, N., and R. Ellouz. 1996. Microbial consortia for the aerobic degradation of aromatic compounds in olive oil mill effluent. J. Indust. Microbiol. 16:155–162.