Introduction

The principal relevance of ligninolytic fungi to the field of bioremediation lies in their ability to degrade aromatic compounds. There are three groups of aromatics that constitute substantial pollutants: polyaromatic hydrocarbons (PAHs), benzene/toluene/ethyl benzene/xylene (BTEX) and the synthetic substituted aromatics typified by the chlorophenols. It may well be that ligninolytic fungi can play a useful role in bioremediation of all three types of pollutant, but the most interest is in degradation of the first and last groups, as BTEX remediation can exploit bacterial populations that promise to be efficient contributors to the process. We will largely be concerned with systems that are of possible direct application to PAH degradation as the halogenated hydrocarbons are degraded by increasingly well-understood biochemical pathways (see Reddy, Gelpke & Gold, 1998; Reddy & Gold, 1999). One of the main difficulties in the development of practical bioremediation processes rests in bringing metabolically active organisms into contact with the pollutant (see Field et al., 1995; Boyle, Wiesner & Richardson, 1998; Head, 1998; Novotny et al., 1999). The secreted enzyme systems of ligninolytic fungi may prove to be a powerful tool for PAH removal, and it is this aspect of their biochemistry to which this chapter is directed.

PAHs are a class of carcinogenic chemical that are formed whenever organic materials are burned; the amount of PAHs in soils coming from atmospheric fall-out have been rising steadily over the twentieth century.

Introduction

Environmental pollutants are a serious concern worldwide because of the hazards they pose to the health of humans and animals. An estimated 80 billion pounds of hazardous organopollutants are produced annually by the chemical, agricultural, oil, paper, textile, aerospace, and other industries in the USA alone (Aust, 1990). Only about 10% of these wastes are believed to be disposed of in an environmentally safe manner (EPA, 1988; Fernando & Aust, 1994). Traditional methods of disposing of hazardous wastes (physical, chemical, and thermal treatments and land filling) have not always been efficacious. It has been estimated that it costs about one trillion dollars to decontaminate toxic waste sites in the USA alone using traditional waste disposal methods (Barr & Aust, 1994). Considering these staggering costs for cleaning up the environment, an alternative, rapid, efficacious and cost-effective method is needed. One method that has become increasingly popular for decontamination of the environment has been bioremediation. The use of indigenous or suitable introduced microorganisms at contamination sites often provides an efficient and economically attractive solution to the pollution problem. One of the early reports indicated that lignin-degrading white rot fungi, as exemplified by Phanerochaete chrysosporium, can degrade an extremely diverse group of environmental pollutants (Bumpus et al., 1985). Since then, there has been intense worldwide research to unravel the potential of white rot fungi in bioremediation. This ability of white rot fungi to degrade a wide spectrum of environmental pollutants sets them apart from many other microbes used in bioremediation.

Introduction

In recent years, recognition of the importance of biological materials as renewable resources for the production of energy and feed, and as important sources of chemical feedstock for the production of different chemicals, has revived interest in the ancient technology of fungal solid-state fermentation (SSF). In the Orient, commercial SSF is still widely practised for miso, saké, soy sauce and tempe production. Agro-industrial byproducts such as lignocellulosic wastes, corn cobs, sugar cane bagasse and coffee pulp, which frequently create serious environmental problems, could potentially be used as low-cost carbohydrate sources for fungal fermentations; these, in turn, would produce biochemical compounds suitable for the food, chemical and pharmaceutical industries.

Fungi are structurally unique organisms that abound in various ecosystems. They are capable of colonizing a wide range of living and dead tissues, including plants, wood and paper products, agricultural plant residues, and live or dead animal tissues. In composting, fungi colonize a mixture of heterogeneous substrates such as municipal solid waste and cattle manure with straw. Many fungi can grow on solid substrates and secrete extracellular enzymes that break down various polymers to molecules that are then reabsorbed by the fungal colony. Consequently any discussion of fungal biodegradation must cover an extraordinary amount of catalytic capability. Exoenzymes derived from filamentous fungi have diverse roles in nature, being involved in the degradation of many types of agricultural matter. Enzymic activities such as oxidation and hydrolysis are involved in the bioconversion of these wastes.

Introduction

With increasing awareness among the general public of the problems of water pollution has come a realization among effluent dischargers that the colour in effluents represents a problem in itself. Colourless effiuents are less visible, attract less attention and cause less concern than coloured effiuents. This is despite the fact that often chromophores may be present in very small amounts and may pose no significant threat to the environment, other than turning a river red or purple! Having said this, there can be significant problems of toxicity associated with some chromophores (Brown & De Vito, 1993) and many coloured effluents contain damaging materials in addition to chromophores. The focus of this contribution is the removal of colour from effluents, and in particular how wood-rotting fungi can be used for this purpose. It is perhaps useful to consider briefly what alternative processes are available before examining the possible roles of fungi.

The main processes used for colour removal are physicochemical and chemical treatments (Laing, 1991; Cooper, 1995) all of which have some drawbacks. Physicochemical treatments include flocculation and coagulation, adsorption, ion exchange, ultrafiltration and reverse osmosis. These processes (apart from expense) have the problem that contaminant chemicals are not destroyed; they are simply removed from effluents and relocated elsewhere – usually disposed of to landfill or by incineration. Chemical processes mainly involve bleaching using chlorine-based chemicals, ozone or peroxides.

Introduction

There is intense interest in utilizing plants to facilitate remediation of contaminated soils because ‘rhizoremediation’ offers a low-cost and ecologically acceptable approach to dissipating pollutants in soils (Anderson, Guthrie & Walton, 1993). The ability of a limited number of plant species, which are normally endemic to naturally metalliferous soils, to hyperac-cumulate metals is being explored with a view to remediating metal-contaminated soils; the process is termed phytoremediation (Cunningham et al., 1996). Phytoremediation as a technology has advantages and disadvantages, but as most hyperaccumulating species that are being explored with a view to commercial exploitation are in the Cruciferae and are generally non-mycorrhizal, these will not be considered in this review. The degradation of organic pollutants in the rhizosphere has also received considerable interest with a view to developing in situ remediation technologies (Anderson et al., 1993). It is here that mycorrhizal associations have to be considered (Donnelly & Fletcher, 1994; Meharg & Cairney, 2000a).

Rhizosphere degradation of organic pollutants

A wide range of organic pollutants are degraded more rapidly in the rhizospheres of most plant species tested than in bulk soils (Anderson et al., 1993). This ‘rhizosphere effect’ varies according to the chemical being degraded, the plant species used and the soil under study. The following explanations are normally put forward to explain enhanced rhizosphere degradation. First, rhizosphere carbon flow greatly stimulates microbial activity in soil surrounding plant roots, and this enhanced microbial activity results in an enhanced pollutant degradation rates.

Introduction

The aim of this chapter is to review the biodegradation of cyanide and its metal complexes by fungi. However, since the degradation of cyanides by bacteria is in many ways similar to that of fungi, bacterial cyanide metabolism will also be considered. There are also many examples of the degradation and utilization of organic cyanides (nitriles) by both bacteria and fungi, although these are outside the scope of this article and will not be examined. For completion, the ability of fungi to produce cyanide (cyanogenesis) will be briefly discussed, as cyanogenic species have the ability to biotransform or biodegrade cyanide. Reviews that cover more specific aspects of microbial cyanide metabolism include Knowles (1976, 1988), Knowles & Bunch (1986), Raybuck (1992), and Dubey & Holmes (1995).

Cyanide chemistry and toxicity

The identification of cyanide as a poison in bitter almonds and cherry laurel leaves dates back to the early Egyptians (Sykes, 1981). Indeed, hydrogen cyanide (HCN) may account for more human deaths throughout history than any other toxin because of its use in executions and large-scale genocide during World War II (Way, 1981).

Hydrogen cyanide is one of the most rapidly acting metabolic inhibitors known, because of its universal inhibition of respiration. By binding to Fe3+ in cytochrome c oxidase, the terminal oxidase of the mitochondrial or bacterial respiratory chain, cyanide inhibits electron transfer to oxygen, and therefore respiration (Stryer, 1988).

Introduction

Processes of natural bioremediation of lignocellulose involve a range of organisms, but predominantly fungi (Hammel, 1997). Laboratory studies on the degradation of lignocellulose, including wood, straw, and cereal grains, have focused mainly on a few fungal species that grow well in the laboratory and can be readily manipulated in liquid culture to express enzymes of academic interest. Our current understanding of the mechanism of lignocellulose degradation stems from such studies. Although some of these enzymes have economic potential in a range of industries, for example pulp and paper manufacture and the detergent industry, it is frequently expensive and uneconomic to use them for bioremediation of pollutants in soils and water columns. In the successful commercial bioremediation processes developed, whole organisms have been used in preference to their isolated enzymes (Lamar & Dietrich, 1992; Bogan & Lamar, 1999; Jerger & Woodhull, 1999).

Most fungi are robust organisms and are generally more tolerant to high concentrations of polluting chemicals than are bacteria, which explains why fungi have been investigated extensively since the mid-1980s for their bioremediation capacities. However, the species investigated have been primarily those studied extensively under laboratory conditions, which may not necessarily represent the ideal organisms for bioremediation. Fungi in little-explored forests of the world, for example tropical forests, may yet prove to have even better bioremediation capabilities than the temperate organisms currently studied, exhibiting more tolerance to temperature and specialist environments.

Introduction

Fungi play a major role in environmental biotechnology. Their morphological, physiological and reproductive strategies make them especially suited for terrestrial habitats. This book is a testament to their multi-faceted role in the biodegradation of natural and xenobiotic compounds and to the major progress that has been made in our ability to use them as agents for the detoxification of hazardous wastes. Nevertheless, the fact remains that most of the successful applications have been performed in laboratory bench-top experiments. Field trials have been plagued by suboptimal results. Physical parameters such as aeration, moisture, nutrient level, pH, temperature and toxic contaminant level interact with living systems in unpredictable ways. Biological parameters such as predation and competition from the resident microbial populations also contribute to the variability of outcomes for in situ bioremediation. The challenge is to create remediation protocols that can be effective despite these numerous uncontrolled variables.

Two major biological strategies have been employed to increase the effectiveness of microbial bioremediation in field trials. The first is the stimulation of the indigenous population, usually through the delivery of a limiting nutrient. This practice is called biostimulation, and successful applications include use in marine oil spills and polycyclic aromatic hydrocarbon (PAH)-contaminated soils (Atlas & Bartha, 1992; Riser-Roberts, 1998). Nitrogen and phosphorus are the most commonly added nutrients (Liebeg & Cutright, 1999).

Introduction

Bacterial leaching of metals (bioleaching, biomining) from mineral resources has a very long historical record (Rossi, 1990; Ehrlich, 1999). Metals have been mobilized from sulfide minerals using processes that involved autotrophic sulfur-oxidizing microorganisms, for example Thiobacillus spp., although the involvement of microorganisms in this process was demonstrated only in the 1920s (Rudolfs & Helbronner, 1922; Waksman & Joffe, 1922). In 1947, Thiobacillus ferrooxidans was identified in acid mine drainage as part of a microbial community that also included several fungi (e.g. Spicaria sp.) (Colmer & Hinkle, 1947). Several industrial processes have been developed based on these findings for the mining of cobalt, copper, nickel, uranium, zinc and gold (Bosecker, 1997; Rawlings, 1997). However, all industrial applications to obtain metals from a series of solid materials depend on the activities of sulfur-oxidizing microorganisms.

Bioleaching is mainly based on three mechanisms. Besides proton-induced metal solubilization and metal reduction or oxidation, metals can also be mobilized from solid materials by ligand-induced metal solubilization. Organic acids from heterotrophic microorganisms represent such ligands. This is particularly important in the biohydrometallurgical treatment of silicate, carbonate and oxide minerals since these materials cannot be directly attacked by sulfur-oxidizing microorganisms. Further developments should enable heterotrophic leaching to be used to extract metals from non-sulfide ores (Ehrlich, 1999). The broad diversity of heterotrophic organisms provides a huge industrial potential that has been hardly investigated.

Introduction

Knowledge of plant-microorganism interactions is of great importance for bioremediation and phytoremediation. A wide variety of microbial populations live in natural and agricultural soils, and in marginal soils contaminated with xenobiotics. Plant roots strongly influence the surrounding environment, producing the so-called ‘rhizosphere effect’ in which microbial populations are qualitatively and quantitatively altered with, reciprocally, their metabolism directly affecting plant biology and the accompanying biota.

Arbuscular mycorrhizal fungi (AMF) belong to the wide spectrum of soil microbiota and are able to improve the growth of the host plant, particularly in soils of low nutritional status or in those modified by human activity. This positive effect can be ascribed to the improvement of nutrient uptake by mycorrhizal colonized plant roots and the increase of soil volume explored for nutrient uptake by the plant, extending from areas in which nutrients have been exhausted to new regions where they are still available. An understanding of the interactions of arbuscular and vesicular-arbuscular mycorrhizas, together with the remaining soil microorganisms naturally associated with plant roots, will provide the basis for development of an important biotechnological tool for bioremediation.

Bioremediation is a managed or spontaneous process in which biological, especially microbiological, catalysis acts on pollutant compounds, thereby reducing or eliminating environmental contamination (Madsen, 1991).

Bioremediation is an expanding area of environmental biotechnology and may simply be considered to be the application of biological processes to the treatment of pollution. The metabolic versatility of microorganisms underpins practically all bioremediation applications and most work to date has concentrated on organic pollutants, although the range of substances which can be transformed or detoxified by microorganisms includes solid and liquid wastes, natural materials and inorganic pollutants such as toxic metals and metalloids. However, the majority of applications developed to date involve bacteria and there is a distinct lack of appreciation of the potential roles, involvement and possibilities of fungi in environmental bioremediation despite clear and growing evidence of their metabolic and morphological versatility. The fundamental importance of fungi in the environment with regard to decomposition and transformation of both organic and inorganic substrates and resultant cycling of elements is of obvious relevance to the treatment of wastes, while the branching, filamentous mode of growth can allow efficient colonization and exploration of, for example, contaminated soil and other solid substrates. This, together with the growing importance of fungi as model systems in eukaryotic cell and molecular biology, physiology and biochemistry, provides the rationale for this work.

The prime objective of this book is to highlight the potential of filamentous fungi in bioremediation, and to discuss the physiology, chemistry and biochemistry of organic and inorganic pollutant transformations. The chapters are written by leading international authorities in their fields and represent the latest and most complete synthesis of this subject area.

Introduction

Energetic compounds have important roles in military and civilian applications, and their production represents a considerable portion of the chemical manufacturing industry. Soils and waters at a significant number of sites worldwide have become contaminated with energetic organonitro compounds as a result of manufacturing and decommissioning of ordnance (Rosenblatt et al, 1991). Kaplan (1990) describes hazardous energetic organonitro compounds as a class of synthetic chemical characterized by the presence of a nitroaromatic, nitrate ester or nitramine functional group or moiety. The relative toxicity, mutagenicity and recalcitrance of these compounds in the environment has led to intensive research for innovative technologies to treat contaminated wastes, soils and waters (Kaplan, 1990, 1992; Rosenblatt et al, 1991).

Technologies have been developed to reduce or remove hazardous energetic organonitro compounds from particular waste streams and from the environment in general. Physical treatment technologies include activated carbon absorption, air stripping, filtration and incineration. Chemical treatment technologies include solvent extraction, surfactant precipitation and neutralization (Kaplan, 1990). Biological treatment technologies include denitrification (Kaplan, 1990), batch and continuous fermentation systems (Funk et al, 1995a,b; Razo-Flores et al, 1997; Lenke et al, 1998) and composting (Isbister et al, 1984; Williams, Ziegen-fuss & Sisk, 1992; Funk et al, 1995b; Emery & Faessler, 1997; Tuomi, Coover & Stroo, 1997; Lenke et al, 1998). A biological approach is often desirable because of its relatively low cost compared with chemical or physical treatment technologies and the innocuous nature of the typical by-products, carbon dioxide and water.

Introduction

Although responsible for saving and improving the quality of human life, pesticides have exerted a significant detrimental effect on the environment and have caused serious health problems, resulting in severe criticism of their use (Hayes, 1986). There is often a fundamental conflict between the need for a sustained level of biological activity of a pesticide in the environment and the requirement that the chemical should be degraded to non-toxic and ecologically safe products (Hill, 1978; Casida & Quistad, 1998). The era of modern synthetic pesticides largely dates from 1939 when the insecticidal properties of 1,1,1-trichloro-2,2-bis(p-chlorophenyl)ethane (DDT) were discovered (Tessier, 1982). Unlike naturally occurring organic compounds, which are readily degraded upon introduction into the environment, some pesticides such as DDT are extremely resistant to biode-gradation by native microflora (Rochkind-Dubinsky, Sayler & Blackburn, 1987a). In most cases, the persistence can be explained by the chemical structure and by the degree of water solubility. In addition, some of these pesticides tend to accumulate in organisms at different trophic levels of the food chain. Chlorinated organic pesticides are one of the major groups of toxic chemicals responsible for environmental contamination and an important potential risk to human health (Kullman & Matsumura, 1996).

The most common pesticides are herbicides, insecticides and fungicides, where herbicides account for nearly 50% of all the pesticides used in developed countries and insecticides account for 75% of all pesticides used in developing countries.

Introduction

Laboratory-based studies have shown that fungi are able to degrade a wide range of organic pollutants (see other chapters) and have great potential for use as inoculants to remediate contaminated soil. However, soil is a heterogeneous environment and it is to be expected that experiments using fungal inocula to remove pollutants will show varying degrees of success. For example, soil environmental conditions such as pH, nutrient and oxygen levels may not be optimal for fungal growth or for activity of the fungal extracellular enzymes involved in pollutant transformation. In addition, results from laboratory studies on fungal transformation of persistent organic pollutants (POPs) carried out under optimal conditions in nutritionally defined liquid media are likely to be different from those obtained in the soil environment. Despite this, fungi have been shown to transform a wide variety of POPs in soil and have been used on a large scale to remediate contaminated sites (Lamar et al., 1994). This chapter will first highlight some important issues faced by researchers when using fungi for soil remediation, provide a critical review of previous work concerning fungal transformation of organic pollutants in soil, and then discuss actual field studies using fungal inocula to remediate contaminated soil. Throughout this chapter ‘pollutant’ refers to persistent organic pollutants only.

Introduction

Fungi are of fundamental importance as decomposer organisms and plant symbionts (mycorrhizas) and can comprise the largest pool of biomass (including other microorganisms and invertebrates) in the soil (Wain-wright, 1988; Metting, 1992). They can be dominant in acidic conditions, where the mobility of toxic metals may be increased (Morley et al., 1996), and this, combined with their explorative filamentous growth habit and high surface area to mass ratio, ensures that fungi are integral bioactive components of major environmental cycling processes for metals and other elements including carbon, nitrogen, sulfur and phosphorus (Gadd & Sayer, 2000). There are examples where fungal isolates from soils with high metal contents exhibit higher metal tolerance than isolates from agricultural soils (Amir & Pineau, 1998), while adaptive and constitutive mechanisms of metal resistance are well known in free-living (Gadd, 1993a; Gadd & Sayer, 2000) and mycorrhizal fungi (Meharg & Cairney, 2000). Metals and their compounds, derivatives and radionuclides, interact with fungi in a variety of ways depending on the metal species, organism and environmental conditions, while fungal metabolism can dramatically influence speciation and, therefore, mobility and toxicity (Gadd, 1993a; Gadd & Sayer, 2000). Antagonistic effects between different metal species may also be a significant phenomenon in free-living (Amir & Pineau, 1998) and symbiotic fungi (Hartley et al., 1997). Solubilization mechanisms, for example complexation with organic acids, other metabolites and siderophores, can mobilize metals into forms available for cellular uptake and leaching from the system (Francis, 1994).

Introduction

Interactions with microorganisms have long been recognized as playing a key role in determining the cycling and ultimate fate of metals in the environment. On the one hand, bioleaching from naturally occurring ores or synthetic sources may result in the release and dispersion of metals while, on the other hand, microbial sorption or accumulation processes concentrate and tend to remove metal species from the surrounding environment. The bioconcentration occasioned by the latter processes may also represent an entry path into the food chain, with potentially fatal consequences for higher organisms.

Microbial metal sorption or accumulation processes may be classified as either dependent or independent of metabolism (Blackwell, Singleton & Tobin, 1995). The former occurs in most, if not all microbial forms, sorption depending on the physicochemical nature of the microbial cell wall. Metal sorption or uptake (typically from the surrounding solution) results from chemical and/or physical binding of metal ions to cell wall functional groups and is, in the main, unchanged if the cells are living, denatured or dead. Metabolism-dependent processes are generally slower and involve active metal transport into and localization within the cell interior (Blackwell & Tobin, 1999). In many instances non-active binding occurs first and it is the initially bound metal that is subsequently transported to the cell interior.

The term biosorption has variously been applied to both the overall process of metal uptake by biological materials and the non-metabolic sorption process.

Introduction

Fungal degradation of monoaromatic compounds has clear implications for bioremediation, and the role of fungi in the removal of these contaminants from the environment has been the subject of extensive study. An understanding of the mechanisms involved in the degradation of benzenoid compounds and elucidation of the catabolic pathways is also important for predicting the recalcitrance of new products in the environment. Furthermore, enzymes catalysing key steps in a catabolic pathway could be used in the design and operation of biosensors for detecting environmental pollutants.

In view of the manifold types of monoaromatic compounds that enter the environment from various sources, this chapter has been confined to coverage of chlorinated monoaromatics and the BTEX group of compounds (benzene, toluene, ethylbenzene and m-, o and p-xylenes). Moreover, since there are already many excellent reviews available, emphasis has been given to the results of research conducted since the early 1990s. The contents cover the sources and distribution of BTEX and chlorinated monoaromatic environmental contaminants, fungal transformation studies including degradation pathways and associated enzymology, and various fungal-based bioremediation strategies employed for contaminant removal.

Sources and distribution of chlorinated monoaromatic and BTEX contaminants in the environment

Monomeric aromatic compounds are widely distributed in the environment as a result of natural synthetic and degradative processes.