Petroleum hydrocarbon contamination in boreal forest soils: a mycorrhizal ecosystems perspective

(a) Mycorrhizosphere bacteria

As mycorrhizal fungi constitute the most significant rhizosphere communities, they have immense potential for interactions with other soil organisms such as bacteria, fungi, protozoa, nematodes, arthropods and mammals, as well as with each other (Fitter & Garbaye, 1994; Read & Perez-Moreno, 2003; Cairney, 2005). The primary factors that influence the composition of associated communities are the quality and quantity of C compounds present, competitive interactions between mycorrhizal fungi and free-living microorganisms for mineral nutrients, and the beneficial, detrimental or neutral impacts of secondary metabolites produced by symbiotic or free-living organisms (Siciliano & Germida, 1998; Cairney & Meharg, 2002). The outcomes of interactions between ECM, ERM and saprotrophic fungal mycelia may include mutual interference of growth (deadlock) or replacement of one taxon with another through competition (Cairney, 2005).

The interactions between ECM/ ERM fungi and the heterotrophic bacterial community are important for accessing mineral nutrients (Burke & Cairney, 1998). Observations that enhanced decomposition of organic compounds occurs in (mycor)rhizosphere soils have been attributed to the greater metabolic activities associated with higher densities of microorganisms (Heinonsalo et al., 2000). Enriched bacterial communities, often arranged as biofilms (organised systems consisting of layers of biologically active cells), have been noted at the surfaces of the ECM fungal mantle and extraradical mycelia, which are the sites of nutrient mobilization, uptake and translocation (Sen, 2003). The exposure of microbial biofilms to organic polymers such as cellulose and proteins appears to drive degradative secondary metabolism; this enables plant and microbial uptake of simple compounds (e.g. sugars, amino acids and mineral nutrients) that are released during the decomposition process (Sen, 2003).

From the germination of fungal propagules in soil to establishment of true symbiosis, mycorrhizal fungi experience a free-living stage during which they interact with bacteria (known as mycorrhizal helper bacteria, MHB) that appear to be beneficial to the colonization process via one or more of several proposed mechanisms (Garbaye, 1994). In axenic culture with nutrient limitation, MHB may act by direct trophic interactions (where bacteria provide C substrates or growth factors to the free-living fungi) or by metabolic detoxification of fungal metabolites (e.g. polyphenols, etc.) (Duponnois & Garbaye, 1990). Bacteria that are active at the time of mycorrhizal formation may facilitate recognition between the plant and mycorrhizal fungus, improve the receptivity of the root for fungal colonization, or stimulate fungal growth, thereby increasing encounters between roots and mycelia (Frey-Klett, Pierrat & Garbaye, 1997). MHB also appear to colonize fungal hyphae and stimulate initial mycorrhizal formation through production of vitamins, amino acids, phytohormones and/ or cell wall hydrolytic enzymes, which may influence germination and growth rates of fungal structures, enhance root development and/or decrease susceptibility to infection (Martin et al., 2000). Shishido et al. (1996) found that three strains of fluorescent pseudomonads enhanced spruce seedling growth through mechanisms unrelated to increased mycorrhizal colonization, but growth promotion of pine by two strains was facilitated by an interaction with mycorrhizas. Mycorrhizal root tips tended to support slightly higher populations of Pseudomonas spp. than non-mycorrhizal root tips and additional colonization sites or altered/ enhanced exudation in the mycorrhizosphere were observed. Frey-Klett et al. (1997) found that high levels of bacterial inoculum (MHB Pseudomonas fluorescens BBc6) in the rhizosphere are not necessary for a helper effect to occur.

Another group of naturally occurring, free-living soil bacteria that colonize roots and enhance plant growth when added to seeds and roots are known as the plant growth promoting rhizobacteria (PGPR) (Chanway & Holl, 1991). PGPR activity has been reported in Azospirillum, Bacillus, Clostridium, Hydrogenophaga, Serratia, Staphylococcus, Streptomyces, and Microbacterium species (Chanway, 1997). Holl & Chanway (1992) found that growth of mycorrhizal pine was stimulated by inoculating the rhizosphere with Bacillus polymyxa strain L6, which appeared to be a function of the size of the bacterial population. Plant growth promotion was not attributed to increased symbiosis by the ECM fungus Wilcoxina, and was also unlikely to be due to N fixation as this Bacillus strain contributed to only 4% of seedling foliar N. Rather, stimulation of pine growth may have been a result of bacterial production of plant growth substances such as indoleacetic acid. Other researchers have suggested that PGPR may, at least in the short term, improve the C supply to mycorrhizas by providing an increased supply of N (fixed from the atmosphere) to the plant (Ingham & Molina, 1991; Martin et al., 2000). Microorganisms may directly stimulate plant growth by providing nutrients (e.g. N, P, S) or growth factors (e.g. auxin, cytokinin, gibberellin), increasing root permeability or inducing plant systemic resistance to pathogens. Indirectly, microorganisms may influence other rhizosphere components that influence plant growth, such as increasing legume or alder root nodule number and size, increasing colonization frequency of mycorrhizal fungi, or suppressing deleterious rhizobacteria (Chanway, 1997).

(b) Plant linkages

Plant communities in northern forest ecosystems are linked below ground via the extensive extraradical mycelial network of mycorrhizal fungi (Dahlberg, 2001; Simard & Durall, 2004). Host-specific fungi form intraspecific plant linkages, whereas fungi with more general host requirements may form interspecific linkages that allow for nutrient and C transfer between different tree species. In a microcosm experiment, radiolabelled C transfer through the soil mycelial network has been demonstrated between Sitka spruce [Picea sitchensis (Bong.) Carr.] and pine species (Pinus contorta Dougl. ex Loud. and P. sylvestris L.) (Finlay & Read, 1986). In the field, Simard et al. (1997) demonstrated bidirectional C transfer between Douglas-fir and paper birch (Betula papyrifera Marsh.) via a common mycelial network, with a significant net gain by the shaded Douglas-fir. Mycorrhizal networks appear to have the capacity to mediate significant N transfer among interconnected plants (Casuarina and Eucalyptus pairs); N gradients (between N-rich donors and N-limited receivers) may drive unidirectional N transfer (He et al., 2005). Similarities in the composition of ECM communities associated with various host species in bioassays and field surveys indicate the potential for linkages between varieties of plant species (Kranabetter et al., 1999; Massicotte et al., 1999).

The coexistence of ECM and ERM plants in boreal forests provides many opportunities for sharing ECM and ERM fungi that link plants and translocate nutrients, although little research on this issue has been conducted (Perotto et al., 2002). Vrålstad, Fossheim & Schumacher (2000) demonstrated that fungal strains derived from ECM morphotype Piceirhiza bicolorata constituted assemblages of very close relatives to ERM type Rhizoscyphus ericae. Similarly, Monreal et al. (1999) showed sequence similarity (ITS2 region) between the ECM fungus Phialophora finlandia and R. ericae. In a resynthesis experiment using 12 R. ericae strains on ECM and ERM hosts, Vrålstad et al. (2002b) showed that genetically close relatives of the ERM fungus R. ericae are true ECM partners with conifer (spruce and pine) and angiosperm (birch) species, but no isolates tested formed both ECMs and ERMs. These studies indicate that ECM and ERM plants may share mycobionts of this species complex (known as the R. ericae aggregate) and, based on ITS phylogeny, the ability to form both ECM and ERM symbioses may have evolved with the R. ericae aggregate. Villarreal-Ruiz, Anderson & Alexander (2004) recently reported the ability of a fungus from the R. ericae aggregate to form simultaneously both ECMs and ERMs in culture with Pinus sylvestris and Vaccinium myrtillus seedlings, respectively, based on rDNA sequencing and microscopy.

Due to the complexity of the molecular mechanisms involved in establishment of a tight (host-specific) symbiosis, the type of fungal associations with different plant hosts may not be of great physiological importance under non-contaminated conditions, but may gain ecological importance under stressed environmental conditions (Perotto et al., 2002). Mycelial linkages may influence fungal and plant ecology by providing a source of fungal inoculum to newly growing roots, allowing the C demands of the mycelium to be met by more than one plant and facilitating the transfer of C and mineral nutrients between neighbouring trees (Jones, Durall & Cairney, 2003). It has been proposed that ECM and ERM fungi may contribute to development of plant communities if the net transfer of C and nutrients is predominantly from a pioneer plant species to a late successional species, but a greater awareness of these processes is important for understanding the interactions between trees and understorey vegetation (Dahlberg, 2001). Kernaghan et al. (2003) demonstrated a positive correlation between ECM fungal richness and overstory host tree richness that was explained by resource heterogeneity in combination with the preference (specificity) of ECM fungi for certain plant hosts. Recently, DeBellis et al. (2006) showed that the distributions of ECM fungi in southern mixed-wood boreal forests are influenced by the relative proportions of host tree species. Conservation of stand diversity should therefore support diverse fungal communities. Whether such communities are essential for ecosystem recovery following PHC contamination is still not known. Regardless, minimizing overstory disruption increases the possibility of preserving the integrated below-ground mycelial network with its associated communities, and maximizes its potential to hasten site recovery.

(a) Decomposition

Heterotrophic bacteria and fungi directly decompose complex carbohydrates (mainly cellulose and lignin) in plant detritus (Wardle, 2002). Cellulose is readily biodegradable as it consists of β(1-4) linkages of D-glucose that form flat, linear chains H-bonded together to create microfibril sheets (Evans & Hedger, 2001). By contrast, lignin is a three-dimensional aromatic polymer consisting of β(0-4) linkages of monomeric units of either cinnamyl alcohol: coumaryl alcohol (grasses), coniferyl alcohol (gymnosperms) or sinaptyl alcohol (angiosperms) that surround the microfibrils and provide rigidity to plant cell walls (Evans & Hedger, 2001). Due to its complex and uniquely heterogeneous structure (i.e. hydrophobicity and thermodynamic stability), lignin is highly resistant to degradation (i.e. recalcitrant) and inhibits decomposition by up to a few years by limiting access of microorganisms or enzymes to substrates (Prescott et al., 2000; Steffen, 2003).

In the early stages of decomposition, soluble compounds and cellulose are rapidly metabolized under conditions where C is available and N is usually limiting (Prescott et al., 2000). Decomposition rates slow over time due to changes in substrate compounds (increased lignin fraction) and succession of microorganisms able to compete for various substrates (Berg, 2000). In the later stages of decomposition, there is a net loss of lignin and N is mineralized from humus (Prescott et al., 2000). Growth may become N-limited in habitats with high C:N ratio substrates, whereas in habitats with low C:N ratio (<30:1) substrates, decomposition of organic matter may result in C limitation (Tiquia et al., 2002). With the loss of cellulose, relative lignin and N concentrations increase and the higher N concentration can slow the decomposition rate. Such slowing may be due to low molecular mass N-containing compounds reacting with lignin residues during humification (Prescott et al., 2000), creating more recalcitrant aromatic compounds; or mineral N may repress synthesis of lignin-degrading enzymes in a wide range of soil organisms (Gallo et al., 2004); or a combination of these (Magill & Aber, 1998). Higher initial N concentration (lower C:N ratio) leads to lower decomposition extent (i.e. lower mass loss) and more stabilized organic matter in forest soils (Berg, 2000). The process of humus formation (i.e. humification) is thought to involve microbial modifications of lignin and condensation of proteins or amino acids into humus precursors, which polymerize into structurally intricate humus molecules (Prescott et al., 2000). Compared to the original plant material, humus is low in carbohydrates (e.g. cellulose and hemicellulose) and high in polyphenolics (e.g. lignin constituents) and immobilized N, which is sparingly available to plants (Prescott et al., 2000).

Although it is generally accepted that mycorrhizas play important roles in decomposition and cycling of C, N and P in ecosystems, details of their functions in nutrient dynamics and regulation of nutrient and energy flows are continuing to develop and be refined (Martin, 2001; Allen et al., 2003). The traditional view of the role of mycorrhizas in obtaining limiting nutrients from forest soils involves fungal exploration for nutrients [e.g. amino acids, ammonium (NH₄⁺), nitrate (NO₃⁻) and inorganic P] that are released during decomposition of plant organic matter by heterotrophic fungal and bacterial communities or that are bound to the soil matrix (e.g. insoluble forms of Al and Ca phosphates) (Martin et al., 2000). However, molecular studies have revealed that some fungal species, previously regarded as decomposers of woody debris (saprotrophs), are both frequent and abundant components of ECM communities (Hibbett et al., 2000; Kõljalg et al., 2000). This, along with other conceptual advances in biocomplexity theory, have led to re-evaluation of how mycorrhizas function within ecosystems and how interactions between multiple species of plants, mycorrhizal fungi and soil saprotrophs regulate community composition and ecosystem processes (Allen et al., 2003).

It has been hypothesized that the distribution of the different mycorrhizal categories is related to specialization for nutrient acquisition in particular environments (Read & Perez-Moreno, 2003). In higher latitude and higher elevation forest ecosystems, where seasonally low temperatures and dry conditions result in very slow rates of decomposition, natural selection may have favoured ECM and ERM symbioses with the capacity to mobilize nutrients from organic material and provide them to plants (Read, 1991; Perotto et al., 2002; Read & Perez-Moreno, 2003). Fungal specialization for N acquisition and utilization may be important for determining community structure, lending strong theoretical support to the idea that ECM diversity increases the effectiveness of nutrient acquisition from different spatial locations and different substrates in soil (Martin et al., 2000; Leake, 2001).

Some ECM fungi appear to be directly involved in nutrient mobilization from organic compounds through production of a wide range of hydrolytic and oxidative enzymes such as polyphenol oxidases (e.g. laccase, catechol oxidase and tyrosinase) and endochitinases (Martin et al., 2000; Burke & Cairney, 2002; Lindahl & Taylor, 2004). Most ECM fungi so far investigated have demonstrated limited phenol-degrading activities, but few have been studied. By contrast, ERM fungi (e.g. R. ericae) appear to have well-developed saprotrophic abilities and degrade most polymeric components (e.g. polysaccharides, lignin, protein, chitin and pectin) of plant and fungal cell walls (Martin et al., 2000). It is widely accepted that enzymatic degradation of organic polymers, through production of an array of hydrolytic enzymes in the extraradical mycelia and translocation of nutrients to the root, is the major benefit of ERM symbioses to plants (Perotto et al., 2002). ERM plants may enhance their exploitation of complex soil substrates by broadening their metabolic capabilities through an association with several fungi endowed with different functional enzymes (Martin et al., 2000). This implies that simultaneous associations with a variety of symbiotic fungi may be an important strategy to broaden the range of functions in the colonization of different substrates (Perotto et al., 2002). Whereas many ECM and ERM fungi appear to have the ability to access N and P directly from organic compounds, the extent to which they contribute directly (via enzymatic catabolism) or indirectly (by influencing soil microbial community structure) to decomposition remains unclear (Cairney & Meharg, 2002; Heinonsalo et al., 2004).

(b) Primary production

Regardless of the mechanisms involved, symbioses with mycorrhizal fungi improve plant nutrient acquisition by facilitating access to organic sources, potentially increasing the uptake of nutrients via extensive growth of the mycelia and circumventing nutrient depletion zones in the soil (Buscot et al., 2000). Nutrients are absorbed across fungal membranes and are either retained by the fungi for biosynthesis and growth, or transported distances of centimeters to meters to the plant roots, where translocation to the host enhances the photosynthetic machinery of the plant (Allen et al., 2003). Plant photosynthetic rates depend on the concentrations of N (for enzymes), P (for ATP and ADP), Fe and Mg (for chlorophyll), internal CO₂ and water (to keep stomata open to fix CO₂) (Buscot et al., 2000). The resulting nutrient sink (root cells) allows for nutrient absorption and translocation through fungal cells that occurs more quickly than diffusion to the roots through soil (Martin et al., 2000). It is thought that, as long as N or P is limiting in the soil environment, plants will support their fungal partners through continued allocation of C (Allen et al., 2003).

By absorbing, assimilating and translocating nutrients (e.g. nitrate, ammonium and amino acids such as glutamate, glutamine and alanine), fungi create a C sink in the mycorrhizal roots (Allen et al., 2003). Heinonsalo et al. (2004) recently reported equivalent ¹⁴C allocations to roots and ECMs in organic and mineral A and B horizons of a podzol, using a Pinus sylvestris mini-rhizotron system. The sink strength controls the rate of photosynthate production and the sugar supply appears to regulate some fungal gene expression (Buscot et al., 2000). Carbon metabolism provides fungal mycelia and plant cells with the energy, reducing power and biomass required for synthesis of various metabolites (e.g. amino acids) required for growth (Martin et al., 2000). Martin et al. (2000) have suggested that differences in ECM morphological features (e.g. abundant emanating hyphae with increased metabolic activity versus few emanating hyphae with decreased activity) may reflect differences in the need for C between taxa. Mycorrhizal fungi acquire most or all C via host photosynthesis and translocation (average 10-20% of net photosynthetic yield), but may also obtain C through assimilation following biodegradation of organic polymers in the soil (Martin et al., 2000; Allen et al., 2003).

The ecological significance of using organic polymers as C sources, which may decrease the need for C from the host plant, is unknown (Martin et al., 2000). A substantial amount of fungal C is allocated to the synthesis of recalcitrant compounds such as chitin (60% of the fungal cell wall) that can persist in the environment for years and increase soil aggregation, stability, C storage and water-holding capacity. It has been estimated that as much as 20% of N in boreal podzolic forest floors may be retained in chitin (β-1,4 linked N-acetylglucosamine units) present in dead and alive fungal mycelia (Lindahl & Taylor, 2004). The remaining C is respired (43-60%), accumulated as fungal storage sugars (e.g. mannitol and trehalose) or lipids, or deposited in the mycorrhizosphere as labile compounds (sugars, amino acids) that support the growth of bacterial communities (Allen et al., 2003; Heinonsalo et al., 2004). The release of microbial communities from C limitation provides the potential for them to play a major role in decomposition and nutrient mobilization (Read & Perez-Moreno, 2003). The C cost for maintaining the external mycelium is unknown, but the extent to which C is allocated from roots to mycelial systems may be intrinsically linked to growth and nutrient-foraging activities of ECM fungi. Using digital autoradiographic techniques, Leake et al. (2001) showed that patterns of C allocation within ECM mycelia are highly dynamic and responsive to changes in niche (caused by spatial variability in resource quality or interactions with other organisms). It has been hypothesized that coexistence among fungi may be explained by the differential partitioning of C resources among fungal species (Allen et al., 2003). The key role of mycorrhizas in C cycling (particularly in the positive feedback loop between plant growth, decomposition and leaf litter quality) may have important consequences for the C gains and losses of ecosystems and thus for the C budget at local, regional and global scales (Read, 1991; Cornelissen et al., 2001; Allen et al., 2003).

(c) Summary

The influences of mycorrhizal fungi on plant populations and communities are not merely the sum of effects on the individuals within populations (Dahlberg, 2001; Koide & Dickie, 2002). Due to their long history and multiple evolutionary events, different plants and fungi bring independent characteristics to the symbiosis, resulting in extensive physiological variation among mycorrhizas (Allen et al., 2003). For example, some fungal species, previously regarded as saprotrophs, are both frequent and abundant components of ECM communities (Hibbett et al., 2000; Kõljalg et al., 2000). The spatial scales within which individual mycelia operate as physically or physiologically integrated entities in nature are also not clear (Cairney, 2005). For example, all groups of ERM fungi reported globally have been found associated with salal from a single site (Berch et al., 2002). Whereas a distinct suite of functions can be assigned to a single mycorrhiza, the many genomic combinations (genetic diversity) of symbionts, environmental heterogeneity and the extensive connectedness of mycorrhizal root systems result in a complex suite of ways that mycorrhizas can function in ecosystems (Cairney, 1999; Allen et al., 2003). For example, certain species of mycorrhizal fungi, rhizosphere organisms and plants may interact such that there is a net immobilization of nutrients, which results in slower rates of decomposition (Allen et al., 2003). Other combinations of organisms (guilds) may not influence the equilibrium of either nutrient cycling or decomposition, or may interact to increase nutrient quality of litter and decomposition rates.

High species richness and abundance may represent ecological adaptation to local environmental heterogeneity and is thought to provide forests with a range of strategies to maintain efficient functioning under an array of environmental conditions (Cairney, 1999; Nannipieri et al., 2003). From a management perspective, a key challenge is to discover if modifications of the environment send mycorrhizal ecosystems in predictable directions. Or, alternatively, is it the combination of fungal and plant species that directs the trajectory? Current evidence suggests a vast range of genetic potential in most mycorrhizal ecosystems ready to respond to changing environmental conditions. Consequently, the hypothesis that environment is predominant in determining the outcome warrants testing.

(a) Chemical toxicity

Petroleum products are complex mixtures that can contain numerous aliphatic (linear and branched chains), alicyclic (unsubstituted and alkyl substituted structures) and aromatic (unsubstituted and alkyl substituted structures with at least one unsaturated ring) compounds (Miller & Herman, 1997; Potter & Simmons, 1998). Numerous methods are available for analysing petroleum hydrocarbon mixtures in environmental media (Weisman, 1998). Natural gases are generally composed of methane, ethane and small amounts of higher molecular mass hydrocarbons whereas most crude oils contain compounds such as paraffins, aromatics, naphthenics and asphaltenes that are present in varying proportions (McGill et al., 1981; Potter & Simmons, 1998). Gasoline typically contains compounds in the nC4 to nC12 range, while diesel compounds are in the nC8 to nC21 range (Potter & Simmons, 1998). Light paraffinic crude oils are dominated by molecules with C numbers less than 16 and consisting mainly of alkanes (paraffins) and cycloalkanes (naphthenes). These chemicals are generally first metabolized by microorganisms (Westlake et al., 1973; Riser-Roberts, 1998). The heavier oils have greater proportions of aromatic hydrocarbons and heterocyclic NSO compounds (i.e. containing N, S, or O) with C numbers usually above 20 (McGill et al., 1981; Delille, Coulon & Pelletier, 2004).

Using chromatographic techniques, crude oils can be resolved into four categories of compounds: asphaltenes, saturates, aromatics and polars (eg. NSO compounds) (McGill et al., 1981; Pollard et al., 1992; Weisman, 1998). Asphaltenes are a mixture of pentane-insoluble, colloidal compounds including polyaromatic and alicyclic molecules with some alkyl substitutes (usually methyl groups) that vary in molecular mass between 500 and several thousand (McGill et al., 1981). The structures of these compounds, particularly those with higher mass, share characteristics with proposed structures of humic acids (Prescott et al., 2000). Saturated and aromatic hydrocarbons (mainly n-alkanes, branched alkanes, mono-, bi-, and polycyclic alkanes and mono-, bi- and polyaromatics) usually account for 75% of the mass of crude oils (McGill et al., 1981; Potter & Simmons, 1998). Up to 25% of the total mass may be n-alkanes, with cyclic hydrocarbons accounting for 30-60% of the total mass. Monocyclic aromatic compounds (e.g. toluene, benzene and xylene) and bicyclic types (e.g. naphthalene, biphenol) represent 1-2%; polycyclic aromatics (usually methylated derivatives of fluorene, phenanthrene, anthracene, chrysene, benzofluorene and pyrene) are present in lower amounts (McGill et al., 1981). The NSO fraction contains polar compounds such as naphthenic acids, mercaptans, thiophenes and pyridines. Most of the N in crude oil is contained in the distillate residue as part of the asphalt and resin fraction and usually accounts for less than 0.2% (rarely exceeds 1%) by mass. The S content varies between 0.3 and 3% whereas the O content usually does not exceed 3% (McGill et al., 1981).

Aliphatic compounds are generally less toxic than aromatics, and toxicity has been found to vary with compound size (McGill et al., 1981; Edwards et al., 1998). The quantity and composition of polycyclic aromatic hydrocarbons (PAHs) are major considerations in the evaluation of toxicity of PHC mixtures (Miller & Herman, 1997). PAHs with four or more benzene rings are known to be genotoxic to humans and other ecological receptors; in general, as relative molecular mass and polarity (i.e. degree of oxidation) of PAHs increase, carcinogenicity also increases and acute toxicity decreases. This is due to the metabolic production of highly reactive electrophilic intermediates that can access biological molecules such as DNA, RNA and proteins and react to form adducts or lesions (Landis & Yu, 1995). Little is known about PHC toxicity to plant and microbial communities in forest soils as the majority of studies have excluded the complex interactions between combinations of chemicals, interacting communities and the soil environment that may exert synergistic, potentiative or antagonistic effects (Landis & Yu, 1995; Evans & Hedger, 2001; Koivula et al., 2004). It is likely that toxicity varies with the type of pollutant, the extent of pollution and the general condition (i.e. extent of obvious signs of stress or disease) of the ecosystem prior to chemical disturbance (Seghers et al., 2003).

In agricultural clay soils, the maximum toxicity of crude oil (indicated by worm survival, seed germination, bacterial bioluminescence and photosynthesis inhibition) was highest immediately after introduction of oil (Chaîneau et al., 2003). An initial decrease in microbial density has often been observed immediately following the addition of PHCs to soil. For example, the addition of 10% (volume/mass) toluene to soil resulted in survival of only about 1% of the indigenous bacteria, which eventually recolonized the soil to reach a high cell density (Huertas et al., 2000). Biotransformations of PHC substrates may also lead to the release of an array of potentially toxic metabolites into the surrounding environment (McGill et al., 1981; Riser-Roberts, 1998). For bacteria, toxicity resulting from PHC contamination has been inferred from decreases in enzyme (hydrogenase and invertase) activity (Suleimanov, Gabbasova & Sitdikov, 2005), although reduced enzyme activities may also result from competition for limiting nutrients following PHC contamination.

The toxicity of non-ionized organic contaminants for microorganisms is primarily due to a nonspecific mode of action that involves partitioning of organic chemicals into the hydrophobic (lipophilic) layer of the cell membrane and disruption of membrane integrity (i.e. increased membrane permeability) (Miller & Herman, 1997). Kirk et al. (2005) suggested that interference with fungal membranes may explain the reduced extraradical hyphal growth of AM fungi (Glomus species) in PHC medium with soil. In general, fungi are considered to be more tolerant of high concentrations of polluting chemicals than bacteria, possibly due to differences in cell wall structure (Blakely et al., 2002). Some yeasts (e.g. Saccharomyces cerevisiae) have been found to alter their membranes (i.e. increase hydrophilicity) to exclude hydrophobic contaminants (Park, Chang & Kim, 1988). It is possible that similar compensation mechanisms occur in other fungi as well. Long-term resistance to PHCs could also be due to the ability of ECM fungi to produce spores that resist environmental stress factors and germinate when the concentration of toxicants associated with PHC contamination has decreased sufficiently over time (Nicolotti & Egli, 1998).

PHCs may directly kill plants on contact, slow their growth, inhibit seed germination, create nutrient-deficient conditions or, at lower concentrations, stimulate plant growth (McGill et al., 1981). Nicolotti & Egli (1998) have suggested that crude oil has a caustic or lethal effect on plants only when it comes into direct contact with tissues and that reduced growth and biomass may be manifestations of changes to soil communities. In general, the taller the trees and the deeper their roots, the greater their tolerance to increased PHC concentrations in soil (Trofimov & Rozanova, 2003).

(b) Soil properties and processes

Blakely et al. (2002) found that creosote impacted soil food webs and decomposition processes more by altering the habitat of microinvertebrates and their prey (i.e. fungi and bacteria) than via direct chemical toxicity. Kirk et al. (2005) suggested that PHCs may interfere with plant-fungus communication by altering root exudation patterns or changing the soil environment such that migration of diffusible chemical signals (e.g. flavonoids, auxins, etc.) is prevented. Many of the major impacts of PHCs on soil biota and plants in forest ecosystems appear to be associated with disturbances to water, nutrient and oxygen supplies related to the hydrophobicity and fluidity of oily products (Tarradellas & Bitton, 1997; Trofimov & Rozanova, 2003).

The disturbance caused by PHC contamination leads to considerable changes in physical and chemical properties that are not typical of unpolluted soils. PHC constituents may be found in mobile form, fixed in the soil pores and fissures, adsorbed on the surface of organic and mineral soil constituents or forming a continuous cover on the soil surface (Trofimov & Rozanova, 2003). Morphological changes in PHC-contaminated podzolic soils exhibit fragmentary patterns resulting from the unevenness of chemical distribution in the soil mass, increased amounts of iron in the upper horizons, and increased amounts of cemented soil aggregates (Trofimov & Rozanova, 2003). The extent of physical movement in the soil profile depends on temperature, PHC viscosity, moisture content, soil structure and soil texture (McGill et al., 1981). Greater lateral spread of PHCs occurs in cold conditions; in hot and dry soil conditions, vertical movement into the water-unsaturated zone may occur more frequently (McGill et al., 1981). In sandy soils, frontal migration of PHCs down the soil profile along the paths of roots and fissures altered the soil profile to a depth of greater than 1 m (Trofimov & Rozanova, 2003). In gray forest soils, heavy fractions of PHCs were retained in the upper layer and filled the largest infiltration, aeration and drainage pores; lighter fractions were largely retained in illuvial horizons and filled fine water retention pores (Suleimanov et al., 2005). This can lead to waterlogging and reducing conditions in the soil profile, both of which inhibit decomposition processes (Trofimov & Rozanova, 2003). Large amounts of oily material in soils may also indirectly increase soil temperature (by 1–10^oC) if there is loss of surface vegetation. In some cases, more damage to the soil may occur due to the high osmotic potential of associated brine water (salinity 40,000 -45,000 μg mL⁻¹) than to the presence of PHCs alone (McGill et al., 1981).

Water insolubility, hydrophobicity and soil sorptive properties increase with increasing size (number of aromatic rings) and complexity (molecule topology or pattern of ring linkages) of chemicals; PAHs with three or more rings tend to be strongly sorbed to the soil (Reilley, Banks & Schwab, 1996; Alexander, 2000; Kanaly & Harayama, 2000; Cerniglia & Sutherland, 2001; Chaîneau et al., 2003). Chemical persistence in soil also depends on several environmental factors, including the type and quality of clay particles (as well as cation exchange capacity), the type and concentration of solutes in surrounding solution, soil organic matter (SOM) content and composition, pH and temperature (Alexander, 1999; Semple et al., 2003). Organic chemicals may be sorbed to and retained by soil particles by adsorption or partitioning. Adsorption entails chemical processes (ion exchange) or, more often, physical forces (H bonding or van der Waals forces) to surfaces of organic polymers or the external surfaces of 1:1 clay minerals and the external and internal surfaces of 2:1 (expanding) clays (Alexander, 1999; Miller & Herman, 1997; Ellerbrock et al., 2005). Sorption to minerals must compete with water and may be very low for nonionized organics in hydrated systems. Sorption to organic solids may occur via physical binding, which concentrates chemicals on outer surfaces or within the pores of a solid. Partitioning of organic chemicals into the SOM is a process of transfer from the bulk state of one phase to the bulk state of another by mechanisms analogous to dissolution and leads to a distribution of molecules within a portion or the entire volume of the organic matter (Alexander, 1999; Chaîneau et al., 2003). The extent of chemical retention in the SOM fraction is directly correlated with the octanol-water partition coefficient of the substance (K_ow, measure of chemical hydrophobicity), the amount of SOM in the solid phase and its degree of oxidation or polarity (Xing, McGill & Dudas, 1994; Alexander, 1999; Wang, Sato & Xing, 2005). In forest soils, the sequestration of organic pollutants in SOM (i.e. sorbed inside soil aggregates or at inactive particle surfaces) may decrease toxicity of chemicals through physical separation from biological receptors, which also decreases substrate bioavailability for enzymatic degradation (Alexander, 2000; Ellerbrock et al., 2005). Toxicity of hydrophobic organic contaminants has been found to be less severe for organisms in soils with high humus content (Salminen & Haimi, 1997).

PHC-polluted soils are characterized by lower values of hygroscopic moisture, hydraulic conductivity and water retention capacity (i.e. wettability) compared to unpolluted soils (Trofimov & Rozanova, 2003; Suleimanov et al., 2005). This is related to the spatial arrangement of hydrophobic components within SOM (Roy & McGill, 2000). Higher molecular mass components and their degradation products remain near the soil surface and form crusts that decrease water availability and limit water and gas exchanges between the soil and the atmosphere. The creation of discrete and continuous water-repellent fronts parallel to the soil surface is also recorded in post-fire forest soils (Certini, 2005). Hydrophobic films on the exterior surfaces of soil aggregates reduce the wettability of the soil and increase structure stability (McGill et al., 1981; Certini, 2005). Many PHC-contaminated soils eventually take up water and remain wet; however, long-term (years) hydrophobicity of crude oil contaminated agricultural soils has been documented in western Canada (Roy, McGill & Rawluk, 1999).

The longer some chemicals remain in soil, the more they appear to resist desorption and biodegradation. Weathered (aged) chemical residues have considerable time to interact with the physical and chemical components of soil. Interactions may entail: (1) sorption, most likely via partitioning; or (2) irreversible incorporation into soil organic matter (via humification) by the catalytic activity of a variety of oxidative enzymes present in the soil matrix (Miller & Herman, 1997; Alexander, 1999). PHC pollution has been found to substantially increase the organic C (humic acid) content of soils (Trofimov & Rozanova, 2003). Humification of PHC constituents is explicit to transformation processes. Covalent bonding between organic chemicals and humic polymers (humin, fulvic acid and humic acid) in soil can form stable linkages to dialkylphthalates, alkanes and fatty acids that are resistant to microbial degradation and are not readily extractable with many organic solvents (McGill et al., 1981; Alexander, 1999). Petroleum residues (as indicated by dichloromethane extraction) are associated with soil organic matter (Roy et al., 2003). Sorption of volatile PHCs from adjacent soil has generated hydrophobicity in soils not directly contaminated with PHCs (Roy & McGill, 2000). They may be either directly incorporated through H bonding of phenolic and benzene carboxylic acids into the molecular structure of soil humic materials or adsorbed to the surface of the molecule (McGill et al., 1981). They are not entirely associated with the humic component, however, because exhaustive extraction with NaOH did not eliminate hydrophobicity of soils (Roy & McGill, 2000). It is not known whether complexes between hazardous organic chemicals and soil humic materials are cleaved in nature to give detectable levels of the original compound, whether these complexes are assimilated by animals and plants, or whether they pose problems of present or future toxicological significance (Alexander, 1999).

In PHC-contaminated soils, the C (energy) supply increases, which promotes metabolic activity on the part of all the microorganisms not directly inhibited by the PHCs and concurrently the C:N ratio tends to increase. The carrying capacity of a soil is the maximum level of microbial activity that can be supported under existing environmental conditions, which depends on the size of the population, availability of O₂ or nutrients, temperature and water availability. The carrying capacity of soils may be exceeded as a result of large inputs of C from PHCs (Miller & Herman, 1997). The intensive growth of PHC-oxidizing microorganisms in response to increased C availability is accompanied by consumption of soil nutrients, resulting in decreased nutrient availability for plants (Xu & Johnson, 1997; Tiquia et al., 2002; Trofimov & Rozanova, 2003). A decrease in available N may also be partly due to the inhibition of nitrification and ammonification processes or from loss of nitrates (McGill et al., 1981; Suleimanov et al., 2005). Some studies have shown that nitrifying bacteria were not found in freshly contaminated soils, but that nitrification processes were eventually regained; stable organic matter and total N content have increased significantly following PHC contamination (McGill et al., 1981). An increase in total N has been attributed to increased atmospheric N₂ fixation during PHC biodegradation (McGill et al., 1981). Thus, as the immobilization of mineral N present in the soil increases, the amount of available N decreases such that organisms benefiting from N₂ fixation, or consortia capable of recycling N from microbial biomass, are the only organisms that can thrive under these conditions (i.e. selection of PHC-tolerant species) (Nicolotti & Egli, 1998). Oxidative degradation may also alter the composition of soil bacterial communities, so that aerobic cellulolytic and proteolytic species decrease and anaerobic N- fixing species increase (Nicolotti & Egli, 1998). Under disturbed water and extreme O₂ limitation, P may be reduced and escape to the atmosphere as hydrogen phosphide (Suleimanov et al., 2005), although this is a very small loss mechanism.

It does not appear that a single addition of PHCs limits microbial communities in the long term. However, several studies indicate that although total microbial numbers tend to increase over time, species richness often decreases, which may or may not have deleterious impacts on ecosystem functions (McGill et al., 1981; Hofman et al., 2004). In general, PHC contamination is expected to lead to an initial loss of richness, followed by rapid proliferation of metabolically competent members of communities inhabiting the new environmental conditions imposed by the chemical contaminants (Gramss, Günther & Fritsche, 1998; Seghers et al., 2003; Díaz, 2004). Some studies have reported drastic reductions in overall ECM biomass and colonization potential in soil following a spill, whereas some fungi appeared resistant to the PHCs and may have benefited from their presence (Nicolotti & Egli, 1998). In greenhouse experiments, spruce seedlings grown in crude-oil-contaminated soils exhibited shifts in ECM community structure in response to increased contaminant concentrations (Nicolotti & Egli, 1998). Nitrogen availability may be a major factor structuring ECM fungal communities; mineral N and foliar nutrient ratios (N:P, P:Al) were found to be excellent predictors of fungal taxonomic richness in organic horizons and organic nitrate availability was a good predictor of their relative abundance (Lilleskov et al., 2002). From studies of impacts of acid rain on soil communities, changes to soil chemical status and functions of the decomposer community have been suggested to lead to imbalances in nutrient cycling and ecosystem productivity (Pennanen et al., 1998).

(a) Bacterial pathways

Because of their metabolic versatility, bacteria are able to obtain energy from virtually every organic compound (Romantschuk et al., 2000; Díaz, 2004). The most common electron acceptor for microbial respiration is O₂ and aerobic processes provide the highest amount of energy to cells. Oxygen is not only the electron acceptor, but also participates in activation of the substrate via oxygenation reactions (Díaz, 2004). Anaerobic (anoxic) conditions are prevalent in aquifers, aquatic sediments and waterlogged soils. Here, biodegradation is carried out by strict anaerobes or facultative organisms using alternative electron acceptors such as nitrate (e.g. denitrifying organisms such as Pseudomonas, Alcaligenes and Flavobacterium), sulphate (e.g. sulphate reducers such as Desulfobacterium), Fe(III) (e.g. ferric iron reducers such as Geobacter), CO₂ (e.g. methanogens such as Methanospirillum) or others such as chlorate, Mn or Cr (McGill et al., 1981; Díaz, 2004). Use of alternative acceptors depends on their availability as well as competition between different microorganisms for electron donors. The energy obtained using Fe(III) or nitrate is almost as efficient as using O₂, but less energy is generated by sulphate reducers and methanogens. Fermentative strains may be energy-limited and restricted to syntrophic existence, requiring other populations to consume the potentially toxic endproducts of fermentation. Photosynthetic organisms use energy from the sun to degrade aromatics anaerobically to acetyl-CoA, which is subsequently used in biosynthetic reactions (Díaz, 2004). In both aerobic and anaerobic degradation, structurally diverse compounds are degraded through many different peripheral pathways to a few intermediates that are further channelled via biochemical pathways (i.e. reactions leading to the formation of Krebs cycle intermediates) to the cell’s central metabolism (Díaz, 2004).

Many microorganisms can use the aliphatic compounds present in PHCs as C sources. The mid-size straight-chain n-alkanes (nC10 to nC18) appear to be metabolized more readily than n-alkanes with shorter or longer chains, and saturated (single C bonds) are degraded more readily than unsaturated (double C bonds) compounds (Miller & Herman, 1997; Delille et al., 2004). The extent and location of hydrocarbon sidechain branching or halogen substitution slows the biodegradation of the compound (Miller & Herman, 1997). The most common aerobic biochemical pathway involves direct incorporation of one atom of O₂ into the alkane by a mixed function oxidase or monooxygenase enzyme, but both O₂ atoms can also be incorporated. In either case, a primary fatty acid is formed that is subjected to consecutive removal of two-carbon fragments (β-oxidation), which are converted to acetyl-CoA. This intermediate enters the Krebs cycle where complete mineralization to CO₂ and H₂O occurs (Miller & Herman, 1997). For alkenes, the first step is attack at the terminal or subterminal methyl group or at the double bond to yield an alcohol or epoxide that can be further oxidized to a primary fatty acid and enter β-oxidation. Degradation of alicyclic hydrocarbons (which are major components of crude oil, constituting 20-67% by volume) is thought to occur primarily via cometabolic reactions to open rings and subsequently cleave linearized products for entry into the Krebs cycle (Miller & Herman, 1997).

Under anaerobic conditions, however, oxygenation of hydrocarbons using O₂ is not possible. Aromatic ring structures may be activated under anaerobic conditions using a reductive rather than oxidative process. There is growing evidence that under anaerobic conditions some microbial communities are able to use O from H₂O or CO₂ for ring cleavage of aromatics or to prepare for ring cleavage of aromatics. For example, Schink, Brune & Schnell (1992) propose H₂O as the O source for ring cleavage during benzoate metabolism by fermenting and denitrifying bacteria, and CO₂ as the O source for carboxylation in preparation for ring cleavage of aniline by a sulphate-reducing bacterium.

Westlake et al. (1973) found that the ability of mixed populations of bacteria to use crude oil as a sole C source depended on the composition and amount of n-saturates, asphaltenes and NSO compounds and also that the aromatic fraction of crude oil was capable of supporting bacterial growth. It seems that PHC biodegradation is more closely related to the intrinsic biodegradability of chemicals (and their bioavailability) than to the particular enzymatic capacities of the microorganisms involved (Chaillan et al., 2004).

Bacteria such as Pseudomonas, Mycobacterium, Rhodococcus, Flavobacterium, Acinetobacter, Arthrobacter, Bacillus and Nocardia are considered the primary degraders of polycyclic aromatic hydrocarbons (PAHs) in soil (Kanaly & Harayama, 2000; Chaillan et al., 2004). Pseudomonas has been the most extensively studied, owing to its ability to degrade so many different contaminants and its ubiquity in soils containing PHCs (McGill et al., 1981; Axelrood et al., 2002a; Delille et al., 2004; Díaz, 2004). Most aerobic peripheral pathways involve oxygenation reactions carried out by monooxygenases and/or hydroxylating dioxygenases that incorporate one or two atom(s) of O₂ into the aromatic ring structure to generate dihydroxy aromatic intermediates (e.g. catechol, protocatechuate, gentisate, hydroxyquinol and hydroquinone) (Miller & Herman, 1997; Siciliano & Germida, 1998; Cerniglia & Sutherland, 2001; Watanabe, 2002; Díaz, 2004). These compounds are the substrates of ring-cleavage enzymes that use molecular oxygen to open the aromatic ring between the two hydroxyl groups (ortho cleavage, catalyzed by intradiol dioxygenases) or proximal to one of the two hydroxyl groups (meta cleavage, catalyzed by extradiol dioxygenases). After several subsequent steps, the linearized products are incorporated into the Krebs cycle. Anaerobic peripheral pathways usually converge to benzoyl-CoA, which is dearomatized by specific multicomponent reductases and consumes ATP (Díaz, 2004).

The initial hydroxylation step is considered to be rate limiting and the enzymes involved generally determine the substrate range of microorganisms, although other factors (e.g. substrate specificity of transcriptional regulators and membrane transporters) may also contribute (Watanabe, 2002). In a recent study of biodegradation by indigenous microbial communities in sub-antarctic soils, PAHs with greater than three rings were generally not degraded (Delille et al., 2004). There is little evidence that microbial growth can be sustained with PAHs with four or more rings as a sole substrate (although they may be degraded by syntrophic cometabolism) and there is very limited information regarding bacterial degradation of PAHs with five or more rings (Reilley et al., 1996; Kanaly & Harayama, 2000). However, bacterial degradation of pyrene (a pericondensed four-ring PAH) has been reported from sediment near a hydrocarbon source, and a Mycobacterium species isolated from that sediment has been shown to degrade pyrene using an inducible enzyme system (Kanaly & Harayama, 2000).

The metabolic flexibility of bacteria is related to their genetic adaptability. For example, Siciliano et al. (2003) found that a substantial decrease of aged PHCs in soil was related to a greater presence of catabolic genes (i.e. alkB, alkane monooxygenase; ndoB, naphthalenedioxygenase; xylE, catechol-2,3-dioxygenase) in bulk and rhizosphere soil. However, it was unclear whether the number of organisms containing these genes increased, or if the number of genetic elements present in the community increased. The genes responsible for aromatic biodegradation pathways are usually arranged in clusters (operons) in mobile genetic elements (e.g. plasmids or transposons). Gene clusters contain catabolic genes (encode enzymes for catabolic pathways), transport genes (responsible for active uptake of the compound) and regulatory genes (adjust expression of the catabolic and transport genes to the presence of the compound to be degraded) (Díaz, 2004). For example, the catabolic genes of the ortho and meta pathways are organized as operons with flanking transposon elements on the TOL (toluene) plasmid (Sarand et al., 1998). This facilitates horizontal transfer of the respective genes and rapid adaptation of microorganisms to the presence of new substrates (Díaz, 2004). Conjugation (transfer of genetic material from one microorganism to another) appears to be important in the dissemination of catabolic genes in the indigenous environment (Sarand et al., 2000; Siciliano et al., 2003).

Depending on chemical structure, contaminant concentration and environmental conditions, the onset of PHC biodegradation generally follows a period of acclimation in which no chemical degradation is evident (Alexander, 1999). Adaptation most commonly occurs by induction of the enzymes necessary for biodegradation, followed by increases in populations of biodegrading organisms (Miller & Herman, 1997). Chronic exposure to PHC substrates (e.g. near natural seepages or in areas where frequent spills occur) results in shorter acclimation periods (due to maintenance of biodegradation pathways within adapted communities) and subsequently increased transformation rates (Miller & Herman, 1997; Alexander, 1999). This pollution-induced community tolerance appears to increase proportionally with increased exposures (Seghers et al., 2003). The end of this period is marked with a rise in respiration and increase in density (varying from slight to several orders of magnitude) that reflects growth of hydrocarbon-degrading populations as well as increased growth of organisms such as protozoa that graze microflora or decompose necrotic tissue (McGill et al., 1981). Subsequent declines in microbial respiration may occur due to complete degradation of labile fractions or to limiting availability of N and P (McGill et al., 1981). High abundance, rapid growth and the ability to transfer genes horizontally allow for rapid microbial adaptation to changes in environmental conditions (Romantschuk et al., 2000; Díaz, 2004).

Genetic changes such as mutations (i.e. appearance of new genotypes) may occur when communities are faced with chemicals that do not have natural chemical analogues (Miller & Herman, 1997). Such events occur at low frequency; however, if new genotypes possess physiological characteristics that provide a selective advantage (e.g. new metabolic capacities), they may multiply (via horizontal gene transfer) within the surviving community (Alexander, 1999). The length of time required for a genetic change or for selection and development of an adapted community is not yet predictable (Miller & Herman, 1997). However, given enough time and favourable environmental conditions, the capacity to degrade almost any organic compound is likely to evolve in or immigrate to a contaminated site (Romantschuk et al., 2000).

(b) Fungal cytochrome P450 and ligninolytic systems

Many fungi (e.g. Aspergillus, Penicillium, Fusarium) isolated from PHC-contaminated soils and cultured on PHC-containing medium have been found to use crude oil as a sole C and energy source (Chaillan et al., 2004). In general, eukaryotic organisms oxidize aromatic compounds to water-soluble products via a cytochrome P450 monooxygenase reaction, incorporating one atom of molecular O₂ into the aromatic ring to form a transient arene oxide and reducing the second atom of O₂ to H₂O. The arene oxide is immediately hydrated by an epoxide hydrolase to yield a trans-dihydrodiol or, alternatively, is non-enzymatically isomerized to form phenols that can conjugate with sulphate, glucuronic acid or glutathione (Miller & Herman, 1997; Cerniglia & Sutherland, 2001). These reactions increase both the water solubility and bioavailability of chemical substrates; soil conditions that favour fungal activity may initially increase the toxicity of the parent chemicals (Reilley et al., 1996). Whereas complete mineralization results in innocuous endproducts (CO₂ and H₂O), partial biodegradation can produce intermediate metabolites with unchanged, reduced or increased chemical toxicity. Toxic chemical intermediates with increased water solubility are of particular concern as this can result in the transport and spread of contaminants through the environment (Miller & Herman, 1997).

The ligninolytic enzyme system of white rot fungi (WRF) has been extensively studied due to structural analogies between lignin and PAHs as metabolic substrates (Scheel et al., 2000). Although lignin is a much larger and more heterogeneous polymer than the fused benzene ring structures of PAHs, it is also hydrophobic and insoluble, thereby posing similar problems for enzyme catalysis (Harvey & Thurston, 2001). WRF degrade lignin using a complex nonspecific enzyme system, often while simultaneously obtaining C from cellulose and hemicellulose (i.e. cometabolism) (Scheel et al., 2000; Steffen, 2003). As with the cytochrome P450 system, the oxidizing enzymes of the ligninolytic system increase the bioavailability, solubility and redox status of the chemical substrates for subsequent metabolism (Harvey & Thurston, 2001). Different fungi appear to possess different combinations of oxidizing enzymes (Harvey & Thurston, 2001).

The initial hydroxylation step of the pathway is accomplished with small, diffusible oxidizing agents (highly reactive radicals) generated by three groups of extracellular enzymes: lignin peroxidases (LiP), manganese peroxidases (MnP) and laccases (Collins & Dobson, 1997; Harvey & Thurston, 2001). LiP (EC 1.11.1.14) and MnP (EC 1.11.1.13) are heme-containing enzymes that function at low pH and catalyze the oxidation of lignin, humic substances and many organopollutants (Schlosser & Höfer, 2002). Both enzymes require H₂O₂, which is generated through fungal glucose oxidase, glyoxal oxidase and arylalcohol oxidase reactions (LiP and MnP) or oxidation of organic acids (MnP only) (Evans & Hedger, 2001). In the white rot basidiomycete Phanerochaete chrysosporium, PAHs with ionization potentials at or below about 7.55eV are substrates for direct one-electron oxidation by LiP, whereas those with ionization potentials above this threshold appear to be acted upon by radical species formed during MnP-dependent lipid peroxidation reactions (Bogan, Schoenike & Lamar, 1996). For LiP, radical cations are produced from one-electron oxidations of non-phenolic compounds, which act as non-specific redox mediators and extend the substrate range and redox capacity of LiP (Harvey & Thurston, 2001). During the catalytic cycle of MnP, the active centre is oxidized by H₂O₂. Reduction of the resting enzyme is achieved by two successive one-electron transfers that oxidize Mn²⁺ to Mn³⁺, which is facilitated by fungal organic acids (e.g. oxalate or malonate) upon chelation of the highly reactive Mn³⁺ state. Schlosser & Höfer (2002) found evidence supporting a physiological role of laccase-catalyzed Mn²⁺ oxidation in providing H₂O₂ for extracellular oxidation reactions and demonstrated a novel type of laccase-MnP cooperation relevant to biodegradation of lignin and organic pollutants.

Laccases (benzendiol:oxygen oxidoreductase, EC 1.10.3.2) are (blue) multicopper enzymes (glycosylated polyphenoloxidases) that are an essential component of a complex nonspecific enzyme system secreted by different kinds of fungi that have been shown to oxidize lignin and various organic contaminants (Schlosser & Höfer, 2002; González et al., 2003; Hoegger et al., 2004). In addition to lignin depolymerization and polyphenol degradation, laccases are thought to be involved in the release of N from insoluble protein-tannin complexes, mycelial pigmentation, humus formation, fruiting body formation and detoxification of phenolic compounds, which protects fungi against soil pollutants and host defense compounds (Kanunfre & Zancan, 1998; Burke & Cairney, 2002; Hoegger et al., 2004). Most reports refer to laccase activity as extracellular, but some WRF may also have intracellular laccases (Burke & Cairney, 2002). Laccases catalyze the reduction of O₂ to H₂O (4 electron reduction without formation of free reduced oxygen species) using a range of phenolic compounds as hydrogen donors (Burke & Cairney, 2002; Schlosser & Höfer, 2002). An electron is removed from the phenolic hydroxyl groups of lignin to form free phenoxy radicals, which are further oxidized to quinines (Hoegger et al., 2004).

(c) Metabolic potential of ECM/ERM fungi

As some ECM and ERM fungi are closely related to WRF, some researchers have suggested that they may have retained some ability to degrade organic substrates, including PHCs (Hibbett et al., 2000; Meharg & Cairney, 2000; Meharg, 2001). Braun-Lüllemann, Huttermann & Majcherczyk (1999) reported on the ability of 16 species (27 strains) of ECM fungi that were isolated from middle European forests to degrade PAHs (1500 ppm of phenanthrene, pyrene, chrysene and benzo[a]pyrene) in pure, liquid culture. The slow but efficient metabolism of benzo[a]pyrene by ECM fungi was comparable to results of experiments with WRF (Braun-Lüllemann et al., 1999). A study in which 58 fungal isolates from different physio-ecological groups were exposed to a range of PAHs showed that all fungal groups could degrade PAHs, but that ECM fungi were 19% as efficient as WRF (Gramss et al., 1999). Similarly, Meharg & Cairney (2000) tested 42 ECM fungal species with several types of persistent organic pollutants and found that 33 species were able to degrade one or more classes of chemicals. Only one of 21 ECM fungal species could not degrade at least one PAH; degrading species seemed to prefer chemicals with four to five rings. In each case, the direct oxidative activities correlated with production of extracellular enzymes that appeared to metabolize aromatic rings (Braun-Lüllemann et al., 1999; Gramss et al., 1999). Green et al. (1999) reported that the ECM fungus Tylospora fibrillosa degraded 4-fluorobiphenyl to significant extents via sequential hydroxylation reactions.

A variety of ECM and ERM fungi are known to produce polyphenol oxidases (e.g. laccase, catechol oxidase and tyrosinase) in culture conditions, but there is little evidence for production of extracellular peroxidases (e.g. LiP and MnP) (Cairney & Burke, 1998; Burke & Cairney, 2002). Timonen & Sen (1998) assayed macerated fungal mycelia from several regions of a microcosm system and found that levels of enzyme activity in the environment were lower for ECM fungi than for saprotrophic fungi, possibly due to avoidance of competition or preferential exploitation of substances of a particular quality. Gramss et al. (1998) reported that most cultured ECM fungi (from sporocarps collected from previously contaminated forest plots and industrial sites) exhibited oxidase activity: high extracellular enzyme activities were found for Lactarius and Russula species and high intracellular enzyme activities were found for Suillus, Hebeloma, Leccinum and Tricholoma species. Donnelly & Entry (1999) found extracellular enzyme activity at the advancing hyphal front of ECM fungi and suggested the possibility of a lack of complete dependence for C on the plant partners.

Few studies have considered mycorrhizal fungi in symbiosis with a plant for the degradation of organic pollutants (Koivula et al., 2004). Meharg, Cairney & Maguire (1997) found that the mineralization rate of ¹⁴C-labelled 2,4-dichlorophenol by ECM fungi (Suillus and Paxillus species) in symbiotic culture with P. sylvestris seedlings was increased by 50% and 250% compared to respective rates of those ECM fungi in axenic culture. As mineralization was extremely slow in vermiculite (i.e. no bacteria present), these data were interpreted to suggest that fungal patch differentiation led to greater enrichment and stability of bacterial communities at the fungal-soil interface. It is unknown how the C contributions of the phytobiont influenced fungal responses or how synergistic or antagonistic interactions between mycorrhizas and other microorganisms altered their ability to mineralize or degrade organic pollutants.

(d) Genetic controls

Ligninolytic enzymes are typically produced by WRF as multiple isoenzymes. This biochemical diversity had been attributed to post-transcriptional modifications of a single gene product, but characterization of several laccase gene families suggests that at least part of this diversity could be due to the multiplicity of laccase genes in fungal genomes (González et al., 2003). For example, Hoegger et al. (2004) found that Coprinopsis cinerea has at least eight different laccase genes within the haploid genome, which is the largest laccase gene family reported so far from a single haploid fungus. Bogan et al. (1996) found that the genome of Phanerochaete chrysosporium contains at least 10 structurally related genes (lipA through lipJ) encoding LiP proteins and at least three MnP genes (mnp1, mnp2, mnp3).

Screening genomes for genes that encode laccases and peroxidases may represent a reliable means of identifying potential enzymatic activities in ECM and ERM fungi (Burke & Cairney, 2002). As DNA sequences for laccase and peroxidase genes are now available for several saprotrophic fungi, opportunities exist to design molecular probes or primers for identification of similar genes and/ or mRNA transcripts in mycorrhizal fungi (D’Souza, Boominathan & Reddy, 1996; Burke & Cairney, 2002). For example, Chen et al. (2003) used laccase gene primers to screen ECM basidiomycetes for laccase-like genes, which were amplified from Lactarius, Russula, Piloderma and Tylospora species. Timonen & Sen (1998) examined gene expression in identified functional components of P. sylvestris mycorrhizal systems and found expression of isozymes (i.e. polyphenol oxidase and acid phosphatase) was increased in hyphal fronts of Paxillus involutus and Suillus bovinus ECM systems as they advanced in the humus. Through gene amplification and sequencing, Chambers et al. (1999) reported evidence for MnP genes and peroxidase activity in cultured Tylospora fibrillosa. Chen et al. (2001) extracted DNA from dried basidiomes or fungal cultures and amplified and sequenced genes for LiP and MnP using primers based on Phanerochaete chrysosporium genes. Although they reported the presence of LiP genes in a broad range of ECM fungal taxa and MnP genes in some ECM fungal taxa (three Atheliaceae taxa), Cairney, Taylor & Burke (2003) recently attempted to repeat these experiments and reported a lack of evidence to support the presence of peroxidase genes in ECM fungi.

Extracellular laccase is constitutively produced in small amounts by several fungi, but enzyme expression is considerably enhanced by a wide range of substances, including a variety of different aromatic compounds (Burke & Cairney, 2002; González et al., 2003). Regulation of laccase production appears to be complex and vary between taxa. For example, in the WRF Pycnoporus cinnabarinus, laccase activity is increased with an increase in C:N ratio whereas in Phanerochaete chrysosporium, laccase activity is repressed by glucose regardless of N content and increased in the presence of cellulose when N is also increased (Burke & Cairney, 2002). From assays of liquid cultures of the ECM fungus Thelephora terrestris, Kanunfre & Zancan (1998) found increased secretion of extracellular laccase with a decreased C:N ratio. At the molecular level, Collins & Dobson (1997) demonstrated that laccase gene (lcc) transcription was activated by copper and nutrient N and that induction occurred at the level of gene transcription in the presence of two aromatic compounds. Chen et al. (2003) reported that some laccase-like genes amplified from Lactarius, Russula, Piloderma, and Tylospora species appeared to be regulated at the transcriptional level, with transcription enhanced by higher N content.