Potential of non-ligninolytic fungi in bioremediation of chlorinated and polycyclic aromatic hydrocarbons.

New Biotechnology Volume 00, Number 00 February 2015

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Research Paper

Potential of non-ligninolytic fungi in bioremediation of chlorinated and polycyclic aromatic hydrocarbons Ernest Marco-Urrea1, Inmaculada Garcıá-Romera2 and Elisabet Aranda2,3 1 2

Department of Chemical Engineering, School of Engineering, Universitat Auto`noma de Barcelona, 08193 Bellaterra, Barcelona, Spain Department of Soil Microbiology and Symbiotic Systems, Estacioń Experimental del Zaidıń, CSIC Granada, Spain

In previous decades, white-rot fungi as bioremediation agents have been the subjects of scientific research due to the potential use of their unspecific oxidative enzymes. However, some non-white-rot fungi, mainly belonging to the Ascomycota and Zygomycota phylum, have demonstrated their potential in the enzymatic transformation of environmental pollutants, thus overcoming some of the limitations observed in white-rot fungi with respect to growth in neutral pH, resistance to adverse conditions and the capacity to surpass autochthonous microorganisms. Despite their presence in so many soil and water environments, little information exists on the enzymatic mechanisms and degradation pathways involved in the transformation of hydrocarbons by these fungi. This review describes the bioremediation potential of non-ligninolytic fungi with respect to chlorinated hydrocarbons and polycyclic aromatic hydrocarbons (PAHs) and also shows known conversion pathways and the prospects for future research.

Introduction Hydrocarbons are usually found in contaminated sites in the form of complex mixtures. In many cases, polycyclic aromatic hydrocarbons (PAHs) and aromatic substitutes such as chloro-organics co-exist and are particularly abundant. Their aromatic nature and the presence of chlorine account for their stability, low water solubility and hydrophobicity and thus their ability to persist under natural conditions. Once released into the environment, biodegradation plays a major role in the destruction of these contaminants. All living organisms catalyze metabolic reactions and contribute to overall biotic activity, with each organism occupying a specific niche and performing a particular function in nature. These metabolic processes also participate in the degradation of xenobiotic compounds, which is the primary mechanism involved in biological

Corresponding author: Aranda, E. ([email protected]) 3

Present address: Department of Microbiology, Water Research Institute, University of Granada, Edificio Fray Luis, Ramoń y Cajal 4, 18004 Granada, Spain. http://dx.doi.org/10.1016/j.nbt.2015.01.005 1871-6784/ß 2015 Elsevier B.V. All rights reserved.

transformation. In this context, bacteria and fungi play a dominant role in bioremediation processes. Fungi may have certain advantages over other microorganisms with respect to bioremediation due to their tolerance to pollutants, their penetration in soil via mycelia and rapid colonization of solid substrates [1]. Most white-rot basidiomycete fungi are capable of producing extracellular ligninolytic enzymes (laccase, manganese peroxidase, versatile peroxidase, lignin peroxidase and dye decolorizing peroxidase) and accessory enzymes (H2O2 generating enzymes and glioxal oxidase), both of which are responsible for the degradation of lignin. The conversion of aromatic compounds by ligninolytic fungi and their enzymatic systems has been studied extensively [2]. Their unspecificity facilitates the oxidation and mineralization of these compounds [3,4]. However, since they mainly grow in compact wood and in the presence of lignocellulosic substrates and also favor acidic conditions (pH 3–5), they are unable to compete with non-ligninolytic fungi in soil. It is therefore doubtful whether ligninolytic fungi are involved in the decomposition of aromatic materials under natural conditions [5]. www.elsevier.com/locate/nbt

Please cite this article in press as: Marco-Urrea, E. et al., Potential of non-ligninolytic fungi in bioremediation of chlorinated and polycyclic aromatic hydrocarbons, New Biotechnol. (2015), http://dx.doi.org/10.1016/j.nbt.2015.01.005

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NBT-747; No of Pages 9 RESEARCH PAPER


Research Paper FIGURE 1

Principal methods used by non-ligninolytic fungi to degrade organic pollutants. P450s and epoxide hydrolases constitute two important groups of phase I oxidation enzymes. They may be involved in the primary intracellular attack on the organic pollutant. The formed metabolites may be secreted in the form of conjugates or may undergo further intracellular catabolism through Phase II reactions which are catalyzed by glutathione S-transferases, NAD(P)H: quinine oxidoreductases and UDP-glucuronosyl transferases, among others. The formed metabolites are more easily secreted into the medium. Adapted from [1].

Metabolic enzymes that catalyze xenobiotic biotransformation and detoxification reactions in eukaryotes are classified as phase I and phase II enzymes. Cytochrome P450 monooxygenases (P450s) and epoxide hydrolases constitute two important phase I oxidation enzyme groups. Phase II reactions are catalyzed by glutathione S-transferases, NAD(P)H quinone oxidoreductases and UDPglucuronosyltransferases, among others [6]. Some of these enzymes are inducible by a variety of xenobiotic compounds [7]. In the case of non-ligninolytic fungi, P450s play an important role in xenobiotic detoxification. However, some extracellular oxidation may occur due to ascomycetes laccases or hydroxyl radical attacks [1] (Fig. 1). Thus, in non-ligninolytic fungi, conversion of pollutants usually requires molecular structures to pass through cell walls which are further converted by membrane cell bound enzymes such as P450s or epoxide hydrolases. In this study, we focus on non-ligninolytic fungi belonging to the ascomycetes – usually known as imperfect fungi and formerly classified as deuteromycetes – and zygomycetes species, both of which are common inhabitants of soils (Table 1).

Biodegradation of selected chloro-organics Polychlorinated biphenyls Polychlorinated biphenyls (PCBs) are a class of compounds containing a biphenyl molecule with multiple chlorines (usually between 2 and 8) that form 209 different congeners (Table 2). Although PCBs tend to sorb to fungal mycelia, few fungal strains are able to degrade them. In general terms, characterization of autochthonous fungal species isolated from PCB contaminated soils shows lower levels of fungal diversity than those typically observed in unpolluted soils [8,9]. Isolated fungal species with greater potential as PCB degraders belong to the Ascomycota 2

phylum. When glucose is omitted from the liquid medium, PCBs are not degraded, indicating that the fungal degradation process occurs cometabolically [10]. It is interesting to note that, unlike white-rot fungi that decreased the extent of PCB degradation through an increase in the number of chlorines, isolated ascomycetes appear to degrade individual PCBs at similar degradation rates regardless of chlorine number in both liquid medium and soil [8–10]. However, degradation of technical mixtures of PCBs (Chlophen A) by the non-ligninolytic fungus Aspergillus niger showed that only the mixture with the lowest total chlorine content (42% chlorine PCBs) was biodegradable, whereas the composition of PCBs with higher chlorination levels (54% and 60% chlorine PCBs) remained untransformed [11]. The only evidence on the degradation pathways of PCBs by nonwhite-rot fungi refers to the transformation of 4-chlorobiphenyl by the non-ligninolytic fungus Paecilomyces lilacinus [12]. Five chlorinated metabolites were identified during 4-chlorobiphenyl degradation by this fungus, including ring fission products [12]. The ability to degrade PCBs, though not to dechlorinate, was also observed by Tigini et al. [8] in relation to six different isolates belonging to the genera Aspergillus, Penicillium, Fusarium and Scedosporium, which did not release chloride ions during 2-chlorobiphenyl, 40 ,40 -dichlorobiphenyl and 20 ,20 ,5,50 -tetrachlorobiphenyl degradation in the liquid medium. In that study, laccase was detected in the culture media of all six fungi tested although the role played by this enzyme in PCB degradation was inconclusive [8].

Polychlorinated dioxins Polychlorinated dioxins, also known as dioxins, are organic compounds with a dibenzo 1,4 dioxin central skeleton (Table 2).

www.elsevier.com/locate/nbt Please cite this article in press as: Marco-Urrea, E. et al., Potential of non-ligninolytic fungi in bioremediation of chlorinated and polycyclic aromatic hydrocarbons, New Biotechnol. (2015), http://dx.doi.org/10.1016/j.nbt.2015.01.005

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TABLE 1

Species

Pollutants

Place of isolation

Carbon source

References

Absidia fusca

Ant

Isolated from an Algerian polluted soil

Ant

[40]

A. cylindrospora

Ant

Constructed wetlands

PAHs

[28]

Alternaria

Phen, Pyr

Agricultural soil

PAHs

[38]

Aspergillus sp.

Lin, BaP

Pesticide manufacturing plant; aged and heavily contaminated oily sludge waste in Mexico

Lin, yeast extract medium

[19,21,50]

A. tamarii

End

Agricultural soil in India

End

[23]

A. niger

PCBs, End, Phen

Collection, pesticide contaminated soil; Sugarcane bagasse

PCBs, Dextrose, Glucose

[11,21,43]

A. terreus

End, Ant, Phe, Pyr, BaP

Soil rhizosphere of root-knot nematodes infested plant; PAH polluted soil; crude oil contaminated soil

Dextrose, PAHs, Sugarcane bagasse

[22,28,36,29]

A. sclerotium

Pyr, BaP

Marine sediment

Glucose

[45]

Botryosphaeria laricina JAS6

End

Agricultural soil in India

End

[23]

Cladosporium herbarum

Phe, Pyr

Agricultural soil

PAHs

[38]

C. cladosporoides

Phe

Sugarcane bagasse

Glucose

[43]

C. oxysporum

End

Egg masses of rot knot nematodes

Dextrose

[22]

Cordyceps sinensis

PCDs

Garden

Glucose

[16]

Conidiobolus

Lin

Litter

Glucose (Malt extract glucose yeast peptone)

[18]

Cunninghamella elegans

Phen, Pyr, BaP,

Collection; Agricultural soils

Dextrose, PAHs

[42,44,38]

Cyclothyrium sp.

Ant, Phe

Estuarine sediment contaminated by PAHs

Dextrose

[41]

Fusarium sp.

PCBs, PCDs, BaP

Contaminated soil; leaves and bark of Pterocarpus macrocarpus in a heavily used roads in Bangkok

Glucose, dextrose

[8,17,49]

F. culmorum

Ant, Phe

Agricultural soil

Ant, Phe

[38]

F. poae

Lind

Soil of a pesticide manufacturing plant

Lind

[19]

F. solani

Lind, Ant, Pyr, BaP

Soil of a pesticide manufacturing plant; contaminated compost by addition of domestic fuel; mangrove sediments

Lind, Ant, Pyr, Glucose

[19,31,33]

F. ventricosum

End

Isolated through an enrichment technique using endosulfan as a carbon and energy source.

End

[24]

Mucor thermo-hyalospora

End

India Collection

Low amount of sucrose

[25]

M. racemosus

Pyr

Agricultural soil

Pyr

[38]

Paecilomyces lilacinus

PCBs

Wood chip piles

1% Glucose

[12]

Penicillium sp.

PCBs, BaP

Historically contaminated soil; aged and heavily contaminated oily sludge waste in Mexico

PCBs, yeast extract medium

[8,50]

P. janczewskii

Phe, Pyr

Agricultural soils

PAHs

[38]

P. glabrum

Phe, Pyr

Sugarcane bagasse; soil contaminated with PAHs

Glucose

[43,47]

P. ochloclorom

Pyr

PAHs contaminated soil from a former gasworks site in Stockholm

Pyr

[48]

Pseudallescheria boydii

PCDs

Denitrifying activated sludge

Glucose

[13,15,14]

Scedosporium

PCBs

Historically contaminated soil

PCBs

[8]

Trichoderma sp.

Phe, BaP

Aged and heavily contaminated oily sludge waste in Mexico; different soil samples from Mexico

Yeast extract medium, PAHs

[50,30]

T. viride T3

BaP

Laboratory collection

Glucose

[31]

www.elsevier.com/locate/nbt 3 Please cite this article in press as: Marco-Urrea, E. et al., Potential of non-ligninolytic fungi in bioremediation of chlorinated and polycyclic aromatic hydrocarbons, New Biotechnol. (2015), http://dx.doi.org/10.1016/j.nbt.2015.01.005

Research Paper

Fungal species studied in relation to chloroorganic and PAH bioremediation. Place of isolation and carbon source used during the conversion process. Anthracene (Ant), benzo[a]pyrene (BaP), endosulfan (End), lindane (Lin), phenanthrene (Phe), polychlorinated biphenyls (PCBs), polychlorinated dioxins (PCDs), pyrene (Pyr)



TABLE 1 (Continued ) Species

Pollutants

Place of isolation

Carbon source

References

T. hamatum

Phe, Pyr

Agricultural soil

Phe, Pyr

[38]

Talaromyces spectabilis

Phe, Pyr

Crude oil contaminated soil

Sugarcane bagasse

[29]

Ulocladium chartarum

Ant

Constructed wetlands

Ant

[28]

TABLE 2

Molecular formula, molecular weight (Mw), chemical structure, health risk, environmental data defined in the Directive 2013/39/EU, Henry Constant, Water Partition Coefficient and regulation of the chlorinated and polycyclic aromatic hydrocarbons discussed in the manuscript Research Paper

Compounds

4

Health riska

Environmental dataa

Ranking position in the CERCLA Priority List of Hazardous Substances (2013)a

Henry constant (Pa m3 mol1)b/ octanol-water partition coefficient (log Pow)c

Regulation (EPA) MCL (mg L1)d

Toxic; Non carcinogenic

Very toxic to aquatic organisms, it may cause long-term adverse effects in the aquatic environment; Hazardous to the environment.

9 (PAHs)

H = 2.0 101 [51] log Pow = 4.45

8.3

Toxic; Specific toxicity to some organs; Non carcinogenic; Non mutagenic


9 (PAHs)

H = 3.63 [52] log Pow = 4.52

0.0062

Toxic; Non genotoxic

Very toxic to aquatic organisms, it may cause long-term adverse effects in the aquatic environment; Bioaccumulation may occur in crustacea, fish, milk, molluscs and algae.

254

H = 7.5 101 [51] log Pow = 4.88

0.96

Toxic; Carcinogenic; May cause reproductive difficulties; Teratogenic


8

H = 1.3 101 [51] log Pow = 6.04

0.0002

Long-term risks both to humans and animals; Skin changes (chloracne); thymus gland problems; immune deficiencies; reproductive or nervous system difficulties; increased risk of cancer

Very toxic to aquatic organisms; it may cause long-term adverse effects in the aquatic environment; Hazardous to the environment.

5

H range from 4 104 to 1.0 101 [53] log Pow range from: 4.33 + 0.4 n + m [50]

0.0005

Toxic; Some of them can be teratogenic; mutagenic; carcinogenic; immunotoxic and hepatotoxic


Tetrachlorodibenzop-dioxin: 201 Pentachlorodibenzop-dioxin: 205 Hexachlorodibenzop-dioxin: 196 Heptachlorodibenzop-dioxin: 147

H range from 5.2 105 to 7.5 [55,56] log Pow range from: 6.02 to 7.35

0.00000003 for 2,3,7,8-TCDD


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Health riska

Environmental dataa

Ranking position in the CERCLA Priority List of Hazardous Substances (2013)a

Henry constant (Pa m3 mol1)b/ octanol-water partition coefficient (log Pow)c

Regulation (EPA) MCL (mg L1)d

Toxic; Neurotoxic; Carcinogenic; Liver or kidney problems


32(2007)

H = a 1.5 b = 2.7 101 g = 3.7 [57] log Pow = 3.61–3.72

0.0002 Phase-out program

Toxic; Nervous system difficulties Kidneys problems Seizures in children


44

H = 2.4 101 [58]

No criteria set Phase-out program

a: log Pow = 4.94–5.64 b: log Pow = 4.87–5.65

a

Data obtained from the U.S. Agency for Toxic Substances and Disease Registry (http://www.atsdr.cdc.gov/substances/index.asp). 2013 CERCLA Priority List of Hazardous Substances. Henry’s Law constant (water solution), H. Octanol/Water Partition Coefficient as log Pow. d Regulations obtained from EPA, United Stated Environmental Protection Agency. Maximum Contaminant Level (MCL) – highest level of a contaminant that is allowed in drinking water. MCLs are set as close to MCLGs as feasible using the best available treatment technology and taking cost into consideration. b c

Evidence of polychlorinated dioxin degradation by non-ligninolytic fungi is limited to the ascomycetes Pseudallescheria boydii and Cordyceps sinensis although the enzymes involved in such transformations are still unknown. P. boydii, which was isolated from denitrifying activated sludge, was capable of degrading a mixture containing tetrachlorinated up to octachlorinated dioxins (an average of 82% for individual isomers) [13]. When cultivated with octachlorodibenzo-p-dioxin, which was dechlorinated by P. boydii to two isomers of heptachlorodibenzo-p-dioxins and trace amounts of 1- or 2-hydroxydibenzo-p-dioxins [13]. The degradation process occurred cometabolically using glucose as a carbon source [14], with an optimum pH for fungal growth of less than 9 [15], which is an advantage when compared with the acidic conditions required for most white-rot fungi [4]. These fungi were successfully applied in the treatment of soils mixed with fly ash and contaminated with dioxins under laboratory conditions [15]. The fungus C. sinensis, isolated from a garden, dechlorinated 2,3,7-trichorodibenzo-p-dioxin and octachlorodibenzo-p-dioxin at similar degradation rates using either glucose or 1,4-dioxane as a carbon source [16]. In contrast to P. boydii, although neither dechlorinated dioxins nor those possessing hydroxyl substituents were detected, less chlorinated catechols and catechols themselves together with cis,cis-muconates were identified as by-products [16]. Finally, the ascomycete Fusarium sp. VSO7 was isolated in order to study its ability to degrade dibenzo-p-dioxin and dibenzofuran. This fungus was subsequently used in a microbial biocatalyst containing other bacterial and fungal dioxin-degrading strains in order to dechlorinate polychlorinated dibenzo-p-dioxins [17]. However, the role played by Fusarium sp. VSO7 in the transformation of chlorinated dioxins was not proven.

Chlorinated insecticides Chlorinated insecticides constitute a diverse group of synthetically produced organic chemicals with a variable number of chlorine atoms (Table 2). Unlike PCBs and chlorinated dioxins, all the nonligninolytic fungi isolates capable of degrading the lindane chlorinated insecticide used this compound as a carbon source. The first evidence of this was shown in a study of the phycomyceteous fungus Conidiobolus 03-1-56 isolated from litter [18]. In that study, lindane was degraded only after the depletion of easily degradable nutrients in the medium, and lignin-modifying enzymes were excreted but not clearly related to lindane degradation by this fungus. In addition, three Fusarium species were found to metabolize lindane. F. poae and F. solani were isolated from soil located around a pesticide manufacturing plant [19]. It is interesting to note that fungal species belonging to the genera Aspergillus and Penicillium also showed an ability to grow in agar plates with lindane as a sole carbon source in this preliminary screening, although Fusarium species were chosen after exhibiting greater growth potential. In both cultures, the growth rates of F. poae and F. solani increased in the presence of lindane up to 100 mg mL1. A further increase in the lindane concentration from 100 to 600 mg mL1 resulted in a gradual decrease in the fungal biomass, which, however, was higher than that in the control [19]. By contrast, the biomass of F. verticilliodes AT-100, isolated from Agave tequilana leaves, did not record a significant decrease in the range of lindane concentration of 0–1000 mg mL1 [20]. The by-products of lindane degradation by F. verticilliodes AT-100 were identified as g-pentachlorocyclohexane and benzoic acid derivatives [20]. The presence of benzoic acid derivatives resulted in a drastic reduction in pH which may inhibit fungal growth. Endosulfan is a cyclodiene organochlorine which has been used as an insecticide all over the world (Table 2), although currently is


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TABLE 2 (Continued ) Compounds


Research Paper

been banned in many countries. Degradation of endosulfan by non-ligninolytic fungi has been reported in relation to the genera Aspergillus [21–23], Fusarium [24], Botryosphaeria [23], Cladosporium [22] and Mucor [25]. Most studies report a drastic decrease in pH during fungal endosulfan degradation, probably due to dehalogenation of endosulfan, resulting in the formation of HCl and/or excretion of organic acids by fungi [24]. One of the major concerns associated with endosulfan remediation is the accumulation of endosulfan sulfate, which is more toxic and persistent than the parent compound [26]. The degradation pathway of endosulfan described in relation to fungi belonging to the Aspergillus, Botryosphaeria and Cladosporium genera shared the formation of the transient product endosulfan sulfate via oxidation. It was further transformed by A. niger ARIFCC 1053 into the less toxic endosulfan diol probably by P450s. Sulfurous acid, glyoxal and protonated formic acid were also detected, which are easily converted to CO2, SO2 and H2O in the environment [21]. B. laricina JAS6, A. terreus, C. oxysporum and A. tamarii JAS9 also produced endosulfan sulfate as a major metabolite, which, however, gradually decreased over time, although the metabolites were not identified [22,23]. By contrast, F. ventricosum, which did not produce endosulfan sulfate as an intermediate, hydrolyzed endosulfan to the less toxic endosulfan diol and endosulfan ether [24]. Neither by-product accumulated in the medium during the incubation period is regarded as a precursor of subsequent metabolic products [24]. Finally, the fungus M. thermo-hyalospora MTCC 1384 had both oxidative and hydrolytic pathways by transforming endosulfan into endosulfan diol and small amounts of endosulfan sulfate [25]. With regard to the enzymatic system involved in endosulfan degradation, a positive correlation between the chlorine released, dehalogenase activity and protein released for A. niger was observed [21], suggesting that the degradation process probably occurs extracellularly.

Biodegradation of selected PAHs PAHs are persistent organic pollutants formed by the fusion of two or more benzene rings (Table 2). Several screening studies have demonstrated that the ability to oxidize PAHs is very common among non-ligninolytic environmental isolates [27–29]. These studies show a particular type of resistance to different amounts of PAHs. For instance, Trichoderma displayed a high level of resistance to large concentrations of phenanthrene and benzo[a]pyrene in solid media in petri dishes [30]. The principal mechanism leading to PAH degradation includes intracellular accumulation of the contaminant and its subsequent transformation, as exemplified by F. solani for pyrene and benzo[a]pyrene degradation [31,32]. However, extracellular enzymes such as laccase are also produced by some of these fungi and cannot be ruled out as potentially responsible for PAH degradation [33]. The pathway of conversion during PAH detoxification by nonligninolytic fungi follows a general pattern that includes the formation of oxidized metabolites such as hydroxy-, dihydroxy-, dihydrodiol- and quinone-derivatives (phase I), which may subsequently conjugate with sulfate-, methyl-, glucose-, xylose- or glucuronic acid groups (phase II). According to different authors [34–37], these fungal metabolites are more soluble and generally less toxic than the parent PAH, and thus the health risk after treatment decreased. However, information about degradation of metabolites 6


such as glucose or sulfate conjugates by non-ligninolytic fungi is scarcely available in terms of kinetics or biodegradability. For instance, Capotorti et al. [36] performed an ecotoxicity experiment with metabolite 9-phenanthrenesulfate obtained from Aspergillus terreus that showed that this metabolite was easily degradable by microbial consortia derived from polluted soil. Also, the studies of Schmidt et al. [38] revealed that anthracene metabolites 9-anthrone and 9,10-anthraquinone are easily mineralized in soil slurries. And, in addition, Meulenber et al. [39] performed a similar experiment with white-rot fungi metabolites showing that sequential breakdown by fungi followed by indigenous bacteria leads to and effective PAH bioremediation process. Apart from examples like these, there is not enough evidence that the increased biodegradability is the general pattern in all cases. More studies are needed to have a theory about biodegradability of PAH’s fungal metabolites [36,38] (Fig. 2).

Anthracene Anthracene, consisting of three fused benzene rings (Table 2), is mostly used in the production of insecticides, artificial dyes and coating materials and has been extensively studied in microbial transformations. The majority of the most efficient genera of fungi capable of metabolizing have been isolated in different environments and belong to the Fusarium, Aspergillus, Penicillium, Trichoderma and Absida genera. Giraud et al. [28] isolated a total of 40 strains from an artificial wetland and showed that most of these strains were able to degrade 0.01 g L1 anthracene at different rates under liquid cultivation conditions. The most efficient strains were Ulocladium chartarum and Absidia cylindrospora (degradation >80%) in a medium where glucose was exhausted [28]. F. solani has been described as using anthracene as the sole source of carbon with the production of anthraquinone, anthrone and 1,2-benzenedicarboxylic acid generated via the pathway of benzophenone 1,1-dephenyl, 2-propanone and phthalic acid [33]. Laccase was detected in the F. solani fungus and may be involved in the transformation of anthracene by using a similar pathway to that described for white-rot fungi. By contrast, the conversion products of A. terreus and Ab. fusca isolated from PAH polluted soil were aryl sulfates and hydroxyl aryl sulfates [36] and anthraquinone and 1,4-dihydroxyanthraquinone [40], respectively, suggesting the involvement of P450s and transferase enzymes.

Phenanthrene Phenanthrene, an isomeric benzenoid hydrocarbon of anthracene, has been used as a model compound as it is the smallest PAH with both a bay region and a K region. This is important as, depending on the enzymes produced by each microorganism, microbiological attacks are region specific [29]. In the case of A. terreus [36] and Cyclothyrium sp. CBS109850 [41], non-ligninolytic fungi convert phenanthrene into hydroxyphenanthrene, sulfate phenanthrene and hydroxysulfate derivatives. The presence of dihydrodiols in these studies indicates the involvement of P450s. The Alternaria sp. Gr174, F. culmorum Gr59, P. janczewskii Gr150, Cunninghamella elegans JS/2, Cladosporium herbarum Gr5 and T. hamatum Gr56 species isolated from agricultural soil metabolized 14C phenanthrene and pyrene under submerged liquid cultivation conditions over a period of 11 days in the absence of other carbon sources. The most efficient species was C. elegans



Research Paper


FIGURE 2

Phenanthrene pathway produced by non-ligninolytic fungi. Adapted from [41] and [42].

which produced up to 23 sulfate-conjugated metabolites, ten of which were dihydrodiols [38]. In addition, Cerniglia et al. [42] have detected the glucoside conjugate phenanthrene-1-O-b-glucose during the phenanthrene metabolism. In solid state fermentation, A. niger, P. glabrum and C. cladosporoides, isolated from sugarcane bagasse, are capable of reducing phenanthrene by 70%, 23% and 54%, respectively [43]. The species A. terreus, Talaromyces spectabilis, and Fusarium sp., isolated from Mayan crude oil-contaminated soil, also showed a high tolerance to a mixture of phenanthrene and pyrene which is a solid-state microcosm system. These strains efficiently remove both PAHs from soil using an alternative carbon source such as sugar cane bagasse, indicating that the process is cometabolic [29].

Pyrene Pyrene, a four-ring high-molecular weight (HMW) PAH (Table 2), has been widely studied and is regarded as recalcitrant, thermodynamically stable and very difficult to degrade. C. elegans metabolizes pyrene to 1-hydroxypyrene, 1-pyreneglucoside, 1hydroxypyrene-6-glucoside, 1-hydroxypyrene-8-glucoside, 1,6 pyrenequinone and 1,8 pyrenequinone [44]. Studies carried out by Schmidt et al. [38] also found sulfate conjugate hydroxyderivatives such as pyrene sulfate and hydroxypyrene sulfate in aqueous solutions. These metabolites are also produced by the marine fungi A. sclerotium and M. racemosus cultivated in submerged liquid after 8 days of cultivation in the presence of 330 mM pyrene [45]. The P. janthinellum SFU 403 [46], P. glabrum TW 9424 [47] and P. ochloclorom [48] species are also able to metabolize pyrene. P. janthinellum pyrene conversion proceeds through 1pyrenol followed by its 1,6 and 1,8 pyrenequinone [46]. P. glabrum

TW 9424 transforms pyrene into the main metabolites 1-methoxypyrene and 1,6-dimetoxypyrene [47]. Finally, although P. ochloclorom uses pyrene as a sole source of carbon, the metabolites formed during the conversion process have not been elucidated [48].

Benzo[a]pyrene Benzopyrene is formed by the fusion of a benzene ring to pyrene (Table 2). Some non-ligninolytic fungi have shown an ability to convert benzo[a]pyrene, which is a highly persistent 5-ring PAH. Fusarium sp. can produce dihydroxydihydro-benzo[a]pyrene and benzo[a]pyrenequinone cometabolically within 30 days of incubation [49]. Machıń-Ramı´rez et al. [50] have shown that isolated fungi have a greater capacity to degrade benzo[a]pyrene than bacteria. In general, Penicillium, Trichoderma and Aspergillus convert different amounts of benzo[a]pyrene under liquid state cultivation conditions in 5 days. The marine fungi A. sclerotium and M. racemosus also convert benzo[a]pyrene to hydroxybenzo[a]pyrene and benzo[a]pyrenyl sulfate metabolites in a rich dextrose medium [45]. The ascomycetes T. viride and F. solani have also been studied in relation to benzo[a]pyrene degradation, indicating a high dependence on the glucose source to achieve high conversion rates and clearly suggesting the existence of cometabolism [31].

Conclusions and future research Most of the non-ligninolytic fungi described in this review were isolated from contaminated sites and, as a result, were easily adapted to soil and usually exhibited tolerance at high concentrations of pollutants. In addition, some non-ligninolytic fungi are able to use contaminants as a carbon source and show optimal growth at neutral pH. All these characteristics are favorable when



Research Paper

compared to white-rot fungi and render non-ligninolytic fungi suitable candidates for bioremediation treatments in the field. However, it is not always clear whether the biodegradation process proceeds through cometabolism or metabolism. In general, bioremediation is the objective as the contaminant provides a carbon or energy source for the fungi, stimulating their growth by consuming the pollutant and avoiding the introduction of additional substrates. Further research is required in order to explore their physiological and biochemical characteristics prior to application. Although major studies have been carried out to explore the degradation capability of Zygomycetes and Ascomycetes, there are few documented examples of degradation pathways, particularly in relation to chloro-organic compounds. In addition, there is not enough information about the formed metabolites and thus more studies are needed to know about their biodegradability. This information is important in order to assess the toxicity and fate of the transformation products. In addition, almost nothing is known about the enzymes involved in the transformation of pollutants by non-ligninolytic fungi. There are no studies using


proteomic techniques that deal with the genome sequencing of environmentally important fungi, the identification of putative biodegradation genes that may be involved in pollutant degradation and the up-regulation of such encoded proteins. Such studies could shed light on the enzymes involved in this type of pollutant biotransformation, thus enabling a more effective application of non-ligninolytic fungi to fungi-assisted bioremediation.

Acknowledgements E. Aranda gratefully acknowledges the Ministry of Science and Innovation in Spain (project AGL2008-572) and MINECO for ‘Ramoń y Cajal’ contract. We also wish to thank Michael O’Shea for proof-reading the document.

Appendix A. Supplementary data Supplementary data associated with this article can be found, in the online version, at http://dx.doi.org/10.1016/j.nbt.2015.01. 005.

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