On-farm biopurification systems: role of white rot fungi in depuration of pesticide-containing wastewaters


Correspondence: Carlos E. Rodríguez-Rodríguez, Centro de Investigación en Contaminación Ambiental, Universidad de Costa Rica, 2060 San José, Costa Rica. Tel.: (506) 2511 8203; fax: (506) 2253 1363; e-mail: carlos.rodriguezrodriguez@ucr.ac.cr


Environmental contamination with pesticides is an undesired consequence of agricultural activities. Biopurification systems (BPS) comprise a novel strategy to degrade pesticides from contaminated wastewaters, consisting of a highly active biological mixture confined in a container or excavation. The design of BPS promotes microbial activity, in particular by white rot fungi (WRF). Due to their physiological features, specifically the production of highly unspecific ligninolytic enzymes and some intracellular enzymatic complexes, WRF show the ability to transform a wide range of organic pollutants. This minireview summarizes the potential participation of WRF in BPS. The first part presents the potential use of WRF in biodegradation of pollutants, particularly pesticides, and includes a brief description of the enzymatic systems involved in their oxidation. The second part presents an outline of BPS, focusing on the elements that influence the participation of WRF in their operation, and includes a summary of the studies regarding the fungal-mediated degradation of pesticides in BPS biomixtures and other solid-phase systems that mimic BPS.


Direct discharge of pesticide-containing wastewaters into soil or natural water sources constitutes a major problem of point-source contamination. Inappropriate handling of pesticides during agricultural use results in risk of contamination in tasks such as the spraying of pesticides on the field, their dilution and pouring in spraying tanks, the discharge of residues and cleaning of those tanks after application, and utilization in postharvest treatment (Castillo et al., 2008; Karas et al., 2011). Several physicochemical or biological systems to reduce the effect of pesticide contamination have been implemented; however, most of them are costly or require high technology and therefore cannot be applied at farm scale. In this respect, the use of biopurification systems (BPS) arises as a promising strategy to mitigate the impact of pesticides on the environment. BPS first appeared under the name ‘biobeds’ and take advantage of the augmented biological activity of a mixture of components (biomixture, including soil and agricultural wastes), enclosed in an excavation or container designed to collect and treat spills of pesticides and their diluted residues (Castillo et al., 2008). Degradation takes place within BPS thanks to the enhanced metabolic activity of organisms naturally present in the components of the biomixture, which also contain the nutrients necessary to support their growth.

Microbial degradation of pesticides is one of the main processes responsible for their environmental decontamination. Although bacteria are of extreme importance in the process, fungi have also a leading role, in particular the white rot fungi (WRF), for which the degradation of a wide range of pollutants is well known (Gao et al., 2010). Consequently, besides the innate fungal microbiota within biomixtures, the bioaugmentation of BPS with WRF constitutes a potential strategy for the decontamination of pesticide-containing wastewaters.

This minireview intends to describe the ability of WRF to biodegrade pesticides in order to set the potential role they play in BPS. An overview of the enzymatic features of WRF and the wide spectrum of organic contaminants they degrade is also included.

White rot fungi

WRF comprise a physiological group of organisms (mostly basidiomycetes) capable of extensively degrading lignin (a heterogeneous polyphenolic polymer) in all lignocellulosic substrates, both from coniferous softwood and from angiosperm hardwoods (Pointing, 2001; Lundell et al., 2010). They are the only organisms able to depolymerize and mineralize all the components of wood (mainly cellulose, hemicelluloses, and lignin). The bleached appearance of the wood when attacked by the fungi gives name to the group (Pointing, 2001).

Lignin is one of the most abundant polymers in nature and has a biological function as component of plant cell walls; due to the huge amount of lignocellulose produced by plants, decomposition of lignin by WRF has a leading role in maintaining the carbon cycle and releasing products to be consumed by other organisms (Lundell et al., 2010). Degradation of lignin occurs due to the effect of fungal extracellular enzymes, commonly known as lignin-modifying enzymes (LMEs) (Martínez et al., 2005). The high nonspecificity and nonselectivity of the mechanism have proven to have an interesting potential for the removal of several environmental organic pollutants, as explained in the next sections.

Among WRF, Phanerochaete chrysosporium has become the model organism in bioremediation research; however, other WRF such as Pleurotus ostreatus, Bjerkandera adusta, Irpex lacteus, and Trametes versicolor are also known to degrade an increasing number of pollutants (Pointing, 2001; Asgher et al., 2008; Rodríguez-Rodríguez et al., 2012a).

Enzymatic machinery of WRF

The ligninolytic system of WRF consists basically of laccases (Lac, E.C. and peroxidases; the latter comprise three major enzymes, lignin peroxidase (LiP, E.C., manganese-dependent peroxidase (MnP, E.C., and versatile peroxidase (VP, E.C., that belong to the class II peroxidases within the superfamily of the nonanimal heme peroxidases (Hofrichter et al., 2010). Laccases employ O2 as electron acceptor, while H2O2 plays this role for peroxidases.

Laccases oxidize, with low specificity, several compounds while reducing molecular oxygen to water, in a catalytic cycle that includes several intermediate electron transfers between a substrate and copper atoms (Nyanhongo et al., 2007). Despite their high redox potential that permits the oxidation of diverse aromatic compounds (phenols, polyphenols, aromatic amines, methoxy phenols among others), it is not high enough to oxidize several nonphenolic compounds. Nonetheless in the presence of molecules that act as electron-transfer mediators, laccases are able to oxidize such compounds (Bourbonnais & Paice, 1990), and some of these laccase mediators are produced during the metabolism of WRF (Asgher et al., 2008).

Peroxidases use H2O2 or organic hydroperoxides as electron-accepting cosubstrates. Their catalytic cycle goes through the formation of a first intermediate (compound I), result of the binding of the H2O2 to the heme group of the enzyme and the release of a water molecule. A second intermediate, compound II, is formed and then the enzyme ends in its resting state by two-one-electron withdrawals from the reducing substrates (Hofrichter et al., 2010). MnP catalyzes a H2O2-dependent oxidation of Mn2+ to Mn3+, and the Mn3+ ions are stabilized by chelation with organic acids. These chelated ions act as diffusible redox mediators and attack diverse molecules, giving MnP a versatile oxidative ability (Hofrichter, 2002).

LiP in the presence of H2O2 catalyzes the oxidation of veratryl alcohol (a natural fungal secondary metabolite that serves as a redox mediator) which then carries out oxidations of nonphenolic aromatic residues in lignin to generate aryl cation radicals. These radicals after several reactions are subsequently mineralized (Pointing, 2001). VP is considered as a hybrid between MnP and LiP, both structurally and catalytically, and therefore has a wider versatility for electron donors (Hofrichter et al., 2010).

The wide spectrum of substrate oxidation capacity of the reactions catalyzed by the LMEs has been shown to play an important role in the transformation of diverse organic pollutants (Pointing, 2001; Asgher et al., 2008). However, they are not the only enzymes involved in bioremediation processes. In this respect, subsequent research showed the degradation of some pollutants in the absence of LMEs (Yadav & Reddy, 1992), leading to the demonstration of the role of the intracellular cytochrome P450 system in the transformation mechanism. Cytochrome P450 is a large family of cysteinato-heme enzymes widely distributed in nature, which participate in the oxidative transformation of endogenous and exogenous molecules through a mechanism consisting in the insertion of an oxygen atom (from molecular oxygen) into a substrate and the subsequent reduction of the second oxygen atom to a water molecule (Meunier et al., 2004). Participation of cytochrome P450 in the oxidation of organopollutants is based on findings such as the reduction in the degradation rates in the presence of cytochrome P450 inhibitors (Doddapaneni & Yadav, 2004) and the induction of cytochrome P450-codifying genes during exposure to several contaminants (Bezalel et al., 1997; Marco-Urrea et al., 2008).

In addition to laccases and class II peroxidases, two more recently described peroxidase groups found in fungi (but not exclusively in WRF), namely aromatic peroxygenases (APOs, EC and dye-decolorizing peroxidases (DyPs, EC, have shown important degradative potential for organopollutants. APOs belong to the superfamily of heme-thiolate peroxidases and have been described in agaric basidiomycetes (Agrocybe aegerita and Coprinellus radians), but there is molecular evidence of their wide occurrence in other fungi (Aranda et al., 2010). These enzymes share catalytic properties with peroxidases, catalases, and cytochrome P450 monooxygenases (Ruiz-Dueñas et al., 2011) and transfer oxygen from peroxides to diverse organic substrates (aliphatic, aromatic, and heterocyclic compounds), resulting in reactions such as hydroxylation or epoxidation of aromatic rings and benzylic compounds, phenol oxidation, sulfoxidation of tricyclic heterocycles, oxidation of pyridine derivatives, and cleavage of esters (Barková et al., 2011). On the other hand, DyPs comprise a new superfamily of highly stable heme-containing peroxidases found in fungi and bacteria, which show a high catalytic versatility based on the oxidation of several dyes, including high redox potential anthraquinone derivatives only hardly oxidized by other peroxidases, oxidation of phenol, and cleavage of carotenoids (Ruiz-Dueñas et al., 2011). The transformation of diverse polycyclic aromatic hydrocarbons (PAHs) by APOs (Aranda et al., 2010) and the bleach of dyes by DyPs (Sugano et al., 2009) highlight the role of these enzymes in the removal of xenobiotics.

Potential use of WRF for removal of pesticides and other organic pollutants

Application of fungi for environmental purposes appeared in the 1980s (Gao et al., 2010). Despite being outnumbered by bacteria application reports in this topic, fungi and in particular WRF present several features that make them of potential use for the depletion of organic pollutants. First, oxidation by LMEs of WRF is a cometabolic process, which, contrary to bacterial mechanisms, does not need the internalization of the contaminant to intracellular compartments and thus permits the attack to larger molecules and compounds of low solubility (Pointing, 2001). This may also reduce intracellular toxicity to parental compounds and transformation products. Second, the LMEs are highly nonspecific and mostly constitutively produced, therefore permitting the oxidation of very structurally different compounds and reducing the need of pre-exposition or adaptation to contaminated sites, respectively. Third, the natural hyphal growth of WRF in solid substrates may enhance the access to sequestered contaminants of low bioavailability. Lastly, nutritional requirements of WRF enable the use of low-cost lignocellulosic wastes as substrates for colonization and inoculum production in solid-phase bioremediation approaches (Rodríguez-Rodríguez et al., 2011).

The nature of organic pollutants so far known to be transformed by WRF is quite wide. However, in many cases, the degradation by WRF has been demonstrated in pure cultures, and reports of successful real applications are much less abundant. Organic contaminants degraded by WRF include polychlorinated biphenyls (Yadav et al., 1995; Novotný et al., 1997; Beaudette et al., 2000), PAHs, and creosote (Lamar et al., 2002; Byss et al., 2008; Borràs et al., 2010), explosives such as TNT and RDX (Bayman et al., 1995; Hawari et al., 1999), and more recently emerging pollutants such as diverse groups of pharmaceuticals (Rodarte-Morales et al., 2011; Jelić et al., 2012; Rodríguez-Rodríguez et al., 2012a), brominated flame retardants (Zhou et al., 2007), and UV filters (Badia-Fabregat et al., 2012). Similarly, effluents from paper industry or containing phenols (like olive oil mill wastewaters) and synthetic dyes have shown processes such as detoxification, dechlorination, dephenolization, and decolourization after treatment with WRF (Dhouib et al., 2006; Selvam et al., 2006; Blánquez et al., 2008).

In the particular case of pesticides, several WRF have shown interesting degrading abilities, and some examples of studies in liquid media are summarized in Table 1; nonetheless, fungal-mediated transformation of pesticides is not restricted to WRF (Pinto et al., 2012). An important number of reports employed P. chrysosporium, but other WRF such as Trametes species and P. ostreatus have been also evaluated. Bumpus & Aust (1987) reported the removal and even mineralization of the insecticide DDT by P. chrysosporium, up to 50% after 30 days with the formation of intermediate metabolites such as DDD, dicofol, DBP, and 2,2-dichloro-1,1-bis(4-chlorophenyl)ethanol. Similarly, Kennedy et al. (1990) demonstrated the removal of aldrin, dieldrin, heptachlor, chlordane, mirex, and lindane with the same fungus. For lindane, several transformation products such as tetrachlorocyclohexane, tetrachlorocyclohexanol, tetrachlorocyclohexenol, and tetrachlorocyclohexene epoxide have been identified after the treatment with P. chrysosporium (Mougin et al., 1996; Singh & Kuhad, 1999). Tetrachlorocyclohexane and tetrachlorocyclohexanol have also been found during the transformation of lindane by other WRF (Phanerochaete sordida, Cyathus bulleri, and Trametes hirsutus), being tetrachlorocyclohexanol a common product in all the transformation processes with these three fungi (Singh & Kuhad, 1999, 2000). For endosulfan, Kullman & Matsumura (1996) studied the degradation by P. chrysosporium and suggested the participation of two divergent metabolic pathways, oxidative and hydrolytic. Kamei et al. (2011) carried out an analogous study using Trametes hirsuta as the degradative strain. In both studies, the transformation products endosulfan diol and endosulfan hydroxyether were identified. Fragoeiro & Magan (2005) achieved the simultaneous degradation of a mixture of simazine, dieldrin, and trifluralin with the same removal patterns by P. chrysosporium and T. versicolor, which only left approximately 20% of simazine untransformed.

Table 1. Selected examples of degradation of pesticides by lignin-degrading fungi or their LMEs in liquid medium
PesticideFungus or enzyme employedRemovalCommentsReference
  1. ND, no degradation; NI, no information available.

BromoxynilLaccase-mediator system48.8 nmol min−1 U−1Oxidative dehalogenation involved in the processTorres-Duarte et al. (2009)
Niclosamide142.0 nmol min−1 U−1
Bromofenoxim166.2 nmol min−1 U−1
Dichlorophen1257.6 nmol min−1 U−1
Lindane Cyathus bulleri ≈97% after 28 daysTetrachlorocyclohexanol as transformation productSingh & Kuhad (2000)
Phanerochaete sordida ≈93% after 28 daysTetrachlorocyclohexanol and tetrachlorocyclohexene as transformation products
Lindane Trametes hirsuta ≈95% after 28 daysTetrachlorocyclohexane and tetrachlorocyclohexanol as transformation products in both casesSingh & Kuhad (1999)
Phanerochaete chrysosporium ≈90% after 28 days
GlyphosateMnP/Laccase-mediator system100%/90% after 24 h; 100% after 4 days (commercial formulation)Aminomethylphosphonic acid as transformation productPizzul et al. (2009)
Pesticide Mix 34 (22 compounds)MnP80–100% each after 6 days
Lindane Phanerochaete chrysosporium 3.0–4.5% mineralization after 14 daysRole of peroxidases was ruled out, and participation of cytochrome P450 was suggested. Tetrachlorocyclohexene epoxide and tetrachlorocyclohexenol as transformation productsMougin et al. (1996)
ImazalilLaccase-mediator systems85%/> 95% after 24 hMethyl 4-hydroxybenzoate and 4-hydroxybenzoic acid were used as mediators. Reduction in toxicity on mouse fibroblast L929 cells when methyl 4-hydroxybenzoate was used as mediatorMaruyama et al. (2007)
Endosulfan Phanerochaete chrysosporium > 90%/> 95% after 50 h, 61% after 95 h (N rich medium)Carbon-deficient, nitrogen-deficient, and nitrogen-rich conditions tested. Endosulfan sulfate, endosulfan diol, endosulfan hydroxyether, and (tentatively) endosulfan dialdehyde as transformation products. Two divergent pathways (hydrolytic and oxidative) were suggested. Role of extracellular enzymes was ruled outKullman & Matsumura (1996)
Aldrin Phanerochaete chrysosporium 36.5% after 30 days0.6% mineralizationKennedy et al. (1990)
Dieldrin15.5% after 30 days0.5% mineralization
Heptachlor26.1% after 30 days0.5% mineralization
Chlordane36.8% after 30 days9.4% mineralization
Lindane53.5% after 30 days23.4% mineralization
Mirex19.8% after 30 days2.0% mineralization
Diphenylamine Pleurotus ostreatus/Phanerochaete chrysosporium/Trametes versicolor 100% after 2 h/100% after 5 days/100% after 2 h Karas et al. (2011)
ortho-phenylphenol100% after 2 h/70% after 15 days/100% after 2 h 
Thiophanate methyl100% after 2 h/95% after 15 days/100% after 2 h 
Chlorpyrifos≈80% after 2 days/≈80% after 2 days/≈80% after 2 days 
Imazalil100% after 10 days/100% after 10 days/100% after 10 days 
Endosulfan Trametes hirsuta > 90% after 14 daysThe fungal strain was selected as it did not accumulate endosulfan sulfate. Several transformation products: (tentatively) endosulfan dimethylene, endosulfan diol, endosulfan ether, endosulfan lactone, (tentatively) endosulfan hydroxyetherKamei et al. (2011)
Endosulfan sulfate70% after 10 days
MetalaxylAgrocybe semiorbicularis, Auricularia auricola, Trametes versicolor, Dichotomitus squalens, Flammulina velupites, Hypholoma fasciculare, Phanerochaete velutina, Pleurotus ostreatus, Stereum hirsutum0–65% after 42 days Bending et al. (2002)
Terbuthylazine31–97% after 42 days 
Atrazine0–86% after 42 days 
Diuron4–99% after 42 days 
Simazine/dieldrin/trifluralin (mixture) Phanerochaete chrysosporium ≈70–85/100/100% after 25 days Fragoeiro & Magan (2005)
Trametes versicolor ≈70–85/100/100% after 25 days 
Parathion Phanerochaete chrysosporium/Pleurotus ostreatus/Bjerkandera adusta 217/169/145 μmol g-dry wt−1 h−1 Jauregui et al. (2003)
Terbufos23/59/17 μmol g-dry wt−1 h−1 
Azinphos-methyl223/106/203 μmol g-dry wt−1 h−1 
Phosmet220/103/45 μmol g-dry wt−1 h−1 
Tribufos84/45/25 μmol g-dry wt−1 h−1 
PhosmetMicrosomal fraction of Pleurotus ostreatus10 μmol mg-prot−1 h−1Phthalimide as transformation product
Terbufos6 μmol mg-prot−1 h−13,3-Dimethyl-2-thiabutane and diethyl phosphorodithioic acid as transformation products
Azinphos-methyl2 μmol mg-prot−1 h−14-Ketobenzotriazine, azinphos-methyl oxon, O,O,S-trimethyl dithiophosphate, 2,3-dihydroxy-1,2,3-benzotriazine-4-one as transformation products
TrichlorfonNI2,2-Dichloro-1-ketone ethylphosphoric acid dimethyl ester as transformation product
MalathionNIButanedioic acid, [2-(dimethoxyphosphothionyl)-4-ethyl ester] and O,O,S-trimethyl dithiophosphate as transformation products
DDT Phanerochaete chrysosporium 50% after 30 daysMineralization demonstrated. DDD, dicofol, DBP, and 2,2-dichloro-1,1-bis(4-chlorophenyl)ethanol as transformation productsBumpus & Aust, 1987

Some reports have demonstrated the transformation only by means of LMEs, including laccase-mediator systems for the oxidative dehalogenation of bromoxynil, niclosamide, bromofenoxim, and dichlorophen (Torres-Duarte et al., 2009), and the reduction in toxicity after the treatment of imazalil (Maruyama et al., 2007). Moreover, the simultaneous removal of the 22 individual components of the commercial Pesticide Mix 34 (20–100% each) and commercial glyphosate was achieved with purified MnP (Pizzul et al., 2009). On the contrary, in other studies, the role of the extracellular enzymes in the transformation was ruled out (endosulfan and lindane), and instead, the participation of cytochrome P450 was suggested (Kullman & Matsumura, 1996; Mougin et al., 1996). The transformation of phosmet, terbufos, azinphos-methyl, trichlorfon, and malathion with the microsomal fraction of P. ostreatus, yielding diverse intermediate metabolites (Jauregui et al., 2003), supports the role of such intracellular enzymatic processes. In summary, transformation of pesticides has been evaluated with diverse WRF. Removal occurs in whole-cell systems, but it is also possible by enzymatic means with LMEs, usually acting together with mediators in the case of laccase, or even with intracellular enzymatic complexes. The action of these enzymes usually takes place in the early stages of the transformation pathway. Degradation may lead to mineralization in some cases, a crucial point to avoid the accumulation of toxic transformation intermediates; however, toxicological assays on the subject have been rather scarce.

On-farm BPS


BPS started in Sweden during the 1990s, under the name of biobeds, as ‘a simple and cheap construction intended to collect and degrade spills of pesticides on farms’ (Castillo et al., 2008), but also leftovers derived from washing of spraying equipment (Torstensson & Castillo, 1997). They have been used since 1993 when the first prototypes were built and described (Torstensson & Castillo, 1997), and today more than 1500 biobeds are operative in Sweden (Karanasios et al., 2012). That original design spread to other countries, where modifications were applied and led to other BPS configurations, named biomass beds in Italy, biofilters in Belgium, Phytobac and Biobac in France, and kept the original name in places such as United Kingdom, Denmark, or Latin America. Nowadays, BPS are established in more than 25 countries worldwide (Coppola & Trevisan, 2012; Karanasios et al., 2012). Features such as low maintenance requirements and low cost are common to BPS; however, the composition and configurations have been adapted to local climatic conditions, material availability, and legislation.


BPS are excavations or containers filled with a matrix of high biological activity, usually called biomixture that consists of soil, lignocellulosic wastes, and a humified organic substrate, all mixed at different ratios (see Fig. 1, Castillo et al., 2008; Karanasios et al., 2012). Components other than the biomixture differ according to the configuration of the BPS, and besides being involved in the retention of the pollutants, the biomixture is the element directly in charge of the biological degradation of the pesticides, and therefore, this review is focused on the characteristics of this component that permit the colonization and activity of microorganisms, in particular fungi.

Figure 1.

Schematic representation (cross-section) of a typical on-farm BPS. A BPS is constructed with the purpose of retaining and degrading spillages of pesticides or pesticide leftovers derived from washing of spraying equipment. A BPS with three different layers is shown in the figure. The composition of the biomixture layer may vary according to the different materials available in a specific region. The components of the original Swedish biomixture are shown. A clay layer may be added as part of the BPS for decreasing water flow and increasing pesticide retention time. A grass layer is placed in the upper part of the BPS. A block barrier is shown surrounding the excavated area for the BPS, but different materials may be used.

Each element of the biomixture has an important role in the efficiency of the BPS. The soil provides pesticide-degrading microorganisms and contributes to the sorption capacity (Castillo et al., 2008). Although the texture of the soil employed seems to have negligible effect on degradation (Fogg et al., 2004), the use of primed soil with known mineralization capacity (due to pre-exposure to the target pesticide) has resulted in enhanced efficiency in biomixtures compared to unexposed soil (Sniegowski et al., 2011).

The lignocellulosic wastes contribute as important colonization substrates for WRF, with the consequent release of a wide range of more easily degradable substrates. Therefore, the lignocellulosic wastes act as important sources of nutrients, transforming the biomixture into a biostimulated system. These materials are added to biomixtures at up to 50% (vol), but lower amounts have been evaluated (Coppola et al., 2007; Vischetti et al., 2008). In the original biobed design, straw was added at 50% (vol) as lignocellulosic substrate (Castillo et al., 2008).

The humified material typically contributes to moisture control, sorption ability, and degradation of pesticides. Early biomixtures employed peat for this purpose, but the implementation of BPS in other regions required the use of other materials, particularly urban or garden compost, due to the unavailability and high cost of peat in most places (Fogg et al., 2003).

Factors affecting the role of fungi in BPS

Due to the intrinsic design of biomixtures, fungi may play a leading role in the process of degradation that takes place within BPS. In this respect, the addition of lignocellulosic materials enhances the production of LMEs and therefore increases the activity due to WRF. The level of colonization by WRF may be modulated by the choice of the other components of the biomixture. As mentioned above, the trend is to use either peat or compost as humified material. Given its acidic nature, peat tends to reduce the pH of the mixture, therefore favoring fungal activity and production of LMEs (Tavares et al., 2006). On the other hand, compost typically has a neutral or basic pH. The use of compost instead of peat usually results in a higher pH in the biomixture. These conditions promote metabolic degradation of pesticides by active microbial communities such as bacteria, while limiting fungal activity. Nonetheless, fungal-mediated transformation of other pollutants has been achieved at neutral pH in solid matrices like sludge, and therefore, the ability of WRF should not be underestimated in those conditions (Rodríguez-Rodríguez et al., 2010, 2012b). Similarly, higher C/N ratios favor WRF activity as they have been regarded to promote the production of LMEs (Eggert et al., 1996). Compared to peat, compost contains lower C content and higher levels of nutrients including N, and the use of the latter results in biomixtures with low C/N ratios that favor degradation by bacteria. On the contrary, the limitation in N content in peat biomixtures enhances the cometabolic degradation of pesticides mediated by WRF. In this regard for instance, better degradation of the herbicide bentazon by P. chrysosporium takes place under low N conditions (Castillo et al., 2000).

Degradation of pesticides by WRF in BPS and other BPS-like configurations

Aforementioned studies regarding pesticide degradation by WRF (Table 1) were performed in liquid medium, mostly under pure culture conditions in order to demonstrate the transformation ability. However, the use of WRF in BPS implies that degradation occurs in solid-state systems, for which less reports are available. Moreover, the degree of participation in the pesticide removal by WRF depends on their ability to colonize and compete against or, on the other hand, to work in cooperation with the indigenous microbiota of the system. The inoculation of exogenous WRF to BPS, that is, using the BPS as a bioaugmented system, seems a promising way of operation. In this respect, recent demonstration of successful colonization and degradation ability of WRF such as T. versicolor in the removal of pre-existent pharmaceuticals in a mixture of wheat straw and solid sewage sludge under nonsterile conditions (Rodríguez-Rodríguez et al., 2012b) supports this potential application. On the other hand, some authors suggest a low competitive potential of WRF in soil and point out the ecophysiological group of the litter-decomposing basidiomycetes (LDB) as an interesting option for bioremediation approaches. As soil–litter layers in forests are their natural habitat, it has been postulated that LDB may be more adapted to compete, survive, and coexist with other soil microorganisms. LDB produce ligninolytic enzymes similar to those produced by WRF (Baldrian & Šnajdr, 2006), and species from genera such as Stropharia, Agrocybe, and Collybia have been successfully tested for the degradation of several organopollutants, including PAHs (Steffen et al., 2002, 2003), even in nonsterile soil (Steffen et al., 2007), TNT (Herre et al., 1998), and synthetic dyes (Baldrian & Šnajdr, 2006). The transformation of the pesticide DDT in defined liquid medium by pure cultures of some of LDB members (Agrocybe and Marasmiellus) by reductive dechlorination and cytochrome P450-mediated oxidation (Suhara et al., 2011) supports a potential success upon their application in BPS.

Table 2 summarizes the reports on the use of WRF for pesticide degradation in analogous BPS-like solid-phase systems. von Wirén-Lehr et al. (2001) studied the effect of P. chrysosporium in the degradation of isoproturon in mini-biobeds with a biomixture of soil–peat–straw and compared it with nonbioaugmented biobeds. They found a decrease of 78% and > 99% after 28 and 100 days, respectively, with the fungus, vs. around 75% in uninoculated biomixture (after 100 days). Similarly, a 91% reduction in isoproturon concentration after 14 days (coincident with MnP activity) with the identification of N-demethylated metabolites (mono- and didesmethyl isoproturon) was also reported (Castillo et al., 2001b). This fungus was employed in the elimination of bentazon in both sterilized and nonsterilized wheat straw–packed bed reactors (Castillo et al., 2000, 2001a). 4-chloro-2-methylphenoxyacetic acid (MCPA) was also removed from the nonsterile reactor. In another matrix, soil with corncob as the substrate, Kennedy et al. (1990) demonstrated the disappearance of aldrin, dieldrin, heptachlor, chlordane, lindane, and mirex by P. chrysosporium (ranging from 15% for mirex to 86% for aldrin, after 60 days), including some mineralization of lindane and chlordane (< 25% in both cases). In a similar approach, removal of several pesticides was evaluated in soil inoculated with P. chrysosporium and T. versicolor pregrown in wood chips, usually resulting in better performance by the former fungus on simazine, trifluralin, and dieldrin under different water potentials (Fragoeiro & Magan, 2008). Similarly, T. versicolor significantly contributed to degrade atrazine only under dry soil conditions (85% vs. 50% in uninoculated soil after 24 weeks) when sawdust was employed as carrier of the fungus (Bastos & Magan, 2009). Trametes versicolor as well as Hypholoma fasciculare and Stereum hirsutum showed high efficiency to degrade diverse pesticides at important extents in sterilized BPS biomixtures consisting of straw–soil compost (Bending et al., 2002), suggesting that compost-based biomixtures may support fungal cometabolic transformation. Degradation ability of lindane by Ganoderma australe was optimized by central composite design in sandy soil supplemented with wheat straw, for which a maximized biodegradation/biomass ratio corresponded to a temperature of 17 °C, moisture 58%, straw content 45%, N content 8 μg g-1 and lindane content 13 μg g-1 (Rigas et al., 2007).

Table 2. Use of WRF in the degradation of pesticide in BPS-like systems and other solid matrices
LindaneSandy soil and wheat straw Ganoderma australe 56.9 μg g−1 (theoretical)Calculated biodegradation/biomass optimaRigas et al. (2007)
BentazonWheat straw Phanerochaete chrysosporium 100% after 3 days Castillo et al. (2000)
Sterile straw in a packed bed reactor105 μg day−1Best results under N-limiting conditions
MCPAUnsterile wheat straw in a packed bed reactor Phanerochaete chrysosporium 65% after 20 days Castillo et al. (2001a)
Bentazon  75% after 20 days 
IsoproturonWheat straw Phanerochaete chrysosporium 91% after 14 daysMonodesmethyl isoproturon and didesmethyl isoproturon identified as transformation productsCastillo et al., 2001b;
AldrinSoil–corncob Phanerochaete chrysosporium 86% after 60 days Kennedy et al. (1990)
Dieldrin  28% after 60 days 
Heptachlor  20% after 60 days 
Chlordane  28% after 60 days15% mineralization
Lindane  35% after 60 days23% mineralization
Mirex  15% after 60 days 
AtrazineClay soil and sawdust as carrier Trametes versicolor 85–90% after 24 weeksContribution of the fungus only significant at driest conditionsBastos & Magan (2009)
MetalaxylSterile biobed mixture: barley straw/top soil/compost (50%:25%:25% w/w) Trametes versicolor 40% after 42 days Bending et al., 2002;
Stereum hirsutum 54% after 42 days  
Terbuthylazine Trametes versicolor 40% after 42 days  
Hypholoma fasciculare 37% after 42 days  
Stereum hirsutum 79% after 42 days  
Atrazine Trametes versicolor 51% after 42 days  
Hypholoma fasciculare 61% after 42 days  
Stereum hirsutum 57% after 42 days  
Diuron Trametes versicolor 52% after 42 days  
Hypholoma fasciculare 16% after 42 days  
Stereum hirsutum 74% after 42 days  
Iprodione Trametes versicolor 58% after 42 days  
Hypholoma fasciculare 47% after 42 days  
Stereum hirsutum 62% after 42 days  
Chlorpyrifos Trametes versicolor 36% after 42 days  
Hypholoma fasciculare 29% after 42 days  
Stereum hirsutum 6% after 42 days  
SimazineSoil–wood chips Phanerochaete chrysosporium 34–48% after 12 weeks Fragoeiro & Magan (2008)
  Trametes versicolor 27–46% after 12 weeks 
Trifluralin  Phanerochaete chrysosporium 0–30% after 12 weeks 
  Trametes versicolor 5–17% after 12 weeks 
Dieldrin  Phanerochaete chrysosporium 40–46% after 12 weeks 
  Trametes versicolor 5–11% after 12 weeks 
IsoproturonMini-biobed, soil/peat/straw (30:20:50% v/v) Phanerochaete chrysosporium 78% after 28 days, > 99% after 100 days von Wirén-Lehr et al. (2001)

A reduction in the toxicity of the pesticides during the treatment in BPS should be of high concern, as degradation processes may release more toxic transformation products than the original compounds. In this respect, several studies have found a reduction in toxicity after treatment with WRF on pollutants such as diclofenac (Marco-Urrea et al., 2010) or mefenamic acid (Hata et al., 2010), while some works report an increase in toxicity caused by the accumulation of toxic degradation products of selected pharmaceutical compounds (Marco-Urrea et al., 2009; Cruz-Morató et al., 2013). Few studies evaluate the changes in toxicity after WRF application on pesticides. They include the reduction in imazalil cytotoxicity on mouse fibroblasts in treatments with some laccase-mediator systems (Maruyama et al., 2007) and the degradation of endosulfan by T. hirsuta without the accumulation of the toxic intermediate endosulfan sulfate (Kamei et al., 2011). Although they were assayed in liquid media, these results are promising of a similar behavior in biomixtures. Moreover, reduction in toxicity has been accomplished in the solid-phase treatment of sludge with T. versicolor, as revealed by tests with Daphnia magna, Vibrio fischeri, and germination of seeds (Rodríguez-Rodríguez et al., 2011).

Concluding remarks

It is clear that WRF play an important role in the degradation of pesticides in BPS. However, little is known about the exact extent of their participation in these systems. Therefore, more studies aiming to describe in depth the interactions between WRF and other organisms should be conducted, in order to better understand the role of the components of the BPS microbial communities in pesticide elimination (Karanasios et al., 2012). Similarly, the evaluation of pilot BPS bioaugmented with WRF would yield important results regarding the efficiency of such microorganisms at higher scale and longer periods, necessary for a proper optimization of the process. Moreover, standardization of the fungal inoculum production with cheap lignocellulosic wastes could be easily adapted to a low-cost technology such as BPS. The simultaneous use of several fungi and the effect on a wide range of pesticides may lead to the application of BPS to other agro-industrial wastewaters, including those containing antibiotics of agricultural use. Further research should also consider the production of metabolites in situ and include the global evaluation of the changes in toxicity, to better estimate the safety of the process.


This work was supported by Vicerrectoría de Investigación of Universidad de Costa Rica (project 802-B2-046) and IAEA (project COS5029).