# State of the art on fungal biodegradation of persistent organic pollutants in soils and innovative strategies for isolating relevant candidate strains

Charbel Elias, Stéphane Ranque, Laure Malleret

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Charbel Elias, Stéphane Ranque, Laure Malleret. State of the art on fungal biodegradation of persistent organic pollutants in soils and innovative strategies for isolating relevant candidate strains. Environmental Technology and Innovation, 2025, 39, pp.104247. 10.1016/j.eti.2025.104247. hal-05443849

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# State of the art on fungal biodegradation of persistent organic pollutants in soils and innovative strategies for isolating relevant candidate strains

Charbel Elias<sup>a,b</sup>, Stéphane Ranque<sup>a</sup>, Laure Malleret<sup>b,*</sup>

## ARTICLE INFO

Keywords:

Agricultural POPs

Bioremediation

Fungi

Industrial POPs

Persistent organic pollutants (POPs)

Soil

Sustainable remediation strategies

## ABSTRACT

The pollution by persistent organic pollutants (POPs) in terrestrial and soil ecosystems pose a significant global threat, due to their long-term stability, widespread dissemination and detrimental effects on human health and ecological systems. In recent years, myco-remediation has emerged as a sustainable and cost-effective restoration technology, offering several advantages over bacteria, given the omnipresence of fungi, their capacity to tolerate and survive harsh environmental conditions, as well as their ability to oxidize a broader range of compounds, particularly those with high molecular weight and increased recalcitrance. This review explores past and recent advancements in fungal bioremediation of soil contaminated with 19 persistent organic pollutants (POPs) listed under the Stockholm Convention, including ten from the original "dirty dozen" and nine of the newly registered POPs. Fungal species which had demonstrated POP degradation capabilities are delved and their advantages, limitations, and degradation pathways are assessed. A comprehensive review of fungal species applied for POP bioremediation is provided, highlighting the most frequently employed strains. The review also highlights the need of agreeing on standardized approaches, including criteria to select truly effective fungal remediation strains as well as appropriate initial POP concentrations. To guide future research, it provides also insights on innovative fungal isolation approaches, and explores advancements in characterization strategies, to uncover promising candidates for soil bioremediation. Innovations in myco-remediation hold great promise for effectively mitigating POPs in soil environments. This review highlights the growing importance of sustainable solutions for persistent organic pollutants management.

## 1. Introduction

The widespread occurrence of anthropogenic chemicals in soils is a major environmental concern, reflecting the pervasive impact of human activities across various environmental compartments. This ubiquitous contamination stems from various sources, including the intensive use of herbicides, pesticides and fertilizers in agriculture, improper disposal of industrial and biomedical waste, oil spills, nuclear waste accidents, and the discharge of industrial sludge and petroleum products (Khatoon et al., 2021). Persistent Organic Pollutants, as defined and addressed by the Stockholm Convention (Stockholm, 2023) include organochlorine pesticides such as

<footnote><sup>a</sup> Aix Marseille University – APHM, IRD, SSA, UMR RITMES, IHU Méditerranée Infection, Marseille, France</footnote>

<footnote><sup>b</sup> Aix Marseille University – CNRS, UMR 7376, Laboratory of Environmental Chemistry, Marseille, France</footnote>

<footnote>* Corresponding author.</footnote>

<footnote>E-mail address: laure.malleret@univ-amu.fr (L. Malleret).</footnote>
dichlorodiphenyltrichloroethane (DDT), endocrine disruptors such as polychlorinated biphenyls (PCBs), polybrominated diphenyl ethers (PBDE), polybrominated biphenyls (PBBs), in addition to by-products of industrial activities including furans and dioxins (Ren et al., 2018; Alharbi et al., 2018). POPs are carbon-based organic compounds characterized by (i) their proven high toxicity, (ii) their capacity to bioaccumulate in living tissues due to their lipophilic properties which consequently leads to biomagnification up in the food chain, (iii) their capacity to highly resist natural degradation processes which result in their extreme persistence in environmental matrices and finally (iv) their capacity to be transported over large geographic distances, which results in their ubiquitous occurrence (Ren et al., 2018). Moreover, their aptitude to strongly adsorb to soil constituents and aggregates, in particular to clay components and organic matter, results in their sorption and sequestration in the soil matrix, thus limiting their bioavailability and exacerbating their recalcitrance and persistence, especially in aged contaminated soils (Singleton, 2001; Ren et al., 2018).

Exposure to POPs results in severe health problems in humans such as diabetes, obesity, reproductive toxicity, embryotoxicity, cardiovascular diseases, hormonal and neurological disruptions (Alharbi et al., 2018). Although individual POPs exert specific health issues, what renders them worrisome is the fact that organisms are generally exposed to numerous POPs simultaneously due to their increased prevalence as well as their association with various activities and sectors (industrial chemicals, solvents, pesticides, and pharmaceuticals) (Supplementary Table 1). Despite the worldwide recognition of this environmental health threat following the Stockholm Convention, along with the implemented efforts aiming to reduce emissions of persistent organic pollutants by restricting their use, prohibiting their manufacture, and limiting their emissions, POPs remain to date a serious menace to biotic communities (Porta and Zumeta, 2002).

Currently, available remediation technologies include physical, chemical, biological, and combined remediation techniques (Shahsavari et al., 2019). Physical remediation essentially comprises thermal treatment, vitrification, soil flushing, soil washing, in-situ capping, composite membrane filtration, excavation and storage, and electrokinetic (EK) techniques (Song et al., 2017; Bradl and Xenidis, 2005). On the other hand, chemical remediation technologies include chemical oxidation or reduction, adsorption allowing entrapment of contaminants, and photocatalysis (Tong and Yuan, 2012; Nguyen et al., 2020). Bioremediation involves the use of biological agents or their metabolic products for the rehabilitation of contaminated environments. Bacteria, fungi, algae, plants, and enzymes are considered key players in bioremediation processes (Zaborowska et al., 2019).

Bioremediation techniques include (i) bioventing which consists of the aeration of soils in order to stimulate aerobic degradation, (ii) biosparging involving the injection of air and in some cases nutrients in order to stimulate the metabolic activity of indigenous microorganisms, (iii) bioaugmentation which is centered on the inoculation of pollutant-degrading microbial consortia into the contaminated site, thus augmenting the indigenous microbial communities by more capable and potent ones, (iv) biostimulation which is based on the stimulation of native and indigenous microbial communities by providing rate-limiting macronutrients (electron donors, phosphorus, nitrogen, oxygen, carbon sources) thus promoting growth conditions and metabolic activities in order to accelerate bioremediation rates, and lastly (v) phytoremediation which harnesses the capacities of living plants and their rhizosphere to contain, remove, and degrade contaminants (Shahsavari et al., 2019; Prenafeta-Boldú et al. 2018; Song et al., 2017). Recently, combined remediation techniques such as EK-microbial, phyto-microbial, washing-microbial, which combine physicochemical technologies with bioremediation in order to improve the bioavailability of POPs for microbial degradation, have shown promising results (Luo et al., 2006; Barba et al., 2021). In comparison to physical and chemical remediation technologies which are tedious, time-consuming, require high energy inputs, and are usually not cost-effective, bioremediation emerges as a sustainable and environmentally friendly method allowing in-situ rehabilitation of impacted soils. Nevertheless, the efficiency of bioremediation itself or in combination with other physicochemical methods is restricted by several factors, particularly the efficacy of the inoculated strains composing the microbial consortium.

Among these microorganisms, fungi stand out as key players in soil environments, accounting for an important part of the soil biomass and participating in vital processes contributing to nutrient and organic matter cycling, biotic interactions, and nutrient channeling via their mycelial networks. Over the past few decades, laboratory-based studies unveiled that fungi are capable of degrading a broad array of organic pollutants, therefore emerging as promising bioremediation agents (Benguenab and Chibani, 2021; Wongputtisin et al., 2021; Ye et al., 2011; Mancera-López et al., 2008; Čvančarová et al., 2012). Mycoremediation offers several advantages in comparison with bacterial bioremediation: prokaryotic bioremediation is limited among others by the fact that degradation of POPs is a cell-associated mechanism mediated by their intracellular enzymes. However, intake of POPs and their internalization is considered a rate-limiting step, therefore making bacterial degradation inefficient for numerous contaminants, particularly those of high molecular weight. In fungi, POPs transformation occurs through the intracellular and/or extracellular pathways. The intracellular transformation mechanism involves the so-called phase I and phase II enzymes. The cell-bound cytochrome P450 monooxygenases along with epoxide hydrolyses, are two major phase I oxidation enzymes exerting a primordial role in the initial attack on POPs through hydroxylation and epoxidation reactions (Harms et al., 2011a). The resulting hydroxylated products are then subjected to phase II transferases that catalyze conjugation reactions resulting in the formation of conjugates from hydroxylated compounds, thus yielding metabolites with increased solubility and decreased toxicity. On the other hand, the extracellular pathway is associated with ligninolytic fungi belonging namely to the white-rot fungi group (WRF) exhibiting the unique ability to secrete LMEs of low substrate specificity supporting their ability to grow on highly recalcitrant lignocellulosic substrates (Harms et al., 2011b). This pathway is of great interest due to structural similarities between lignin polymers and aromatic pollutants, in addition to the low substrate specificity of LMEs (Supplementary Figure 1). Hence, the use of fungi equipped with a well-developed extracellular enzymatic arsenal of low substrate specificity, renders them capable of targeting a wide range of pollutants, while diffusing effectively to the highly immobile pollutants occurring in their peripheral environment (Benguenab and Chibani, 2021). These enzymatic attacks degrade pollutants into smaller compounds which make them available for intracellular metabolism for both fungi and bacteria. Moreover, fungal mycelial networks characterized by their high surface area to volume ratio, enable exploration of large volumes of
contaminated matrices through their capacity to translocate pollutants over significant distances, while additionally acting as "highways" allowing over-distance transport of pollutant-degrading bacteria (Li et al., 2020).

While existing reviews on fungal remediation of POPs have provided valuable insights, they typically focus on one specific class (Sharma et al., 2018), a couple (Marco-Urrea, García-Romera, and Aranda, 2015) or on selected groups of POPs (León-Santiesteban and Rodríguez-Vázquez, 2017; Grgas et al., 2023). However, the fact that POPs share common properties means that fungal strains need to be compared with each other. In addition, the advantage of reviewing several POPs in a common work enable to highlight which have been little studied and are in need of research, and which strains are used systematically or widely. Hence, the present review strives for comprehensiveness by aiming at encompassing all POPs for which a sufficient body of research and work exists on fungal degradation. In this review, we aim at providing a comprehensive and updated inventory of fungal species capable of degrading an array of selected persistent organic pollutants (POPs). We primarily focus on 19 pollutants categorized into three major POP classes of agricultural, industrial, and by-product origin. These include ten POPs from the Stockholm Convention's initial "dirty dozen" and nine recently added POPs. We delve into the fungal degradation mechanisms for these selected POPs, presenting data on culture conditions, incubation times, degradation percentages, and identified degradation products and intermediates. Additionally, we address emerging POPs for which limited research exists. Finally, the review explores strategies for isolating and characterizing promising fungal strains for soil bioremediation of POPs.

## 2. Methodology

Twenty persistent organic pollutants were selected for in-depth reviewing based on several key criteria. First, all chosen POPs are officially listed under the Stockholm Convention, encompassing both the original "dirty dozen" and subsequently listed "new POPs" ("Listing of POPs in the Stockholm Convention (n.d.)"). Second, each chosen POP must also satisfy at least one of the following criteria. The first criteria was i) a demonstrated mycoremediation potential, where the given POP has been extensively studied in the context of mycoremediation, with evidence of degradation by more than one fungal species. These criteria highlight persistent organic pollutants where mycoremediation shows promise. The second criteria is linked to ii) a public health priority, with the POP being listed under the ATSDR's Substance Priority List (SPL) as a result of its frequency, toxicity, and potential for human exposure at National Priorities List (NPL) sites, which in turn emphasizes POPs that pose potential health risks. This review also briefly explores knowledge gaps in fungal remediation for remaining and emerging POPs. Consequently, this review specifically focuses on 19 POPs, both individually and within the context of the broader POP class. These 19 POPs can be categorized as follows: agricultural pesticides (aldrin, dieldrin, heptachlor, chlordane, endosulfan, endrin, dichloro-diphenyl-trichloroethane, chlordcone, lindane, pentachlorophenol, hexachlorobenzene) and industrial POPs and by-products (polybrominated diphenyl ethers or PBDEs, polychlorinated biphenyls or PCBs, polychlorinated dibenzo-p-dioxins or PCDDs and polychlorinated dibenzo-p-furans or PCDFs).

Concerning literature search and selection, scientific publications were selected from databases like ScienceDirect and search engines such as Google Scholar (for a wider range of academic papers) to ensure better coverage of the topic. The literature review was conducted using the following keywords: the given Persistent Organic Pollutant (the name of each POP was searched independently, e. g. "Endosulfan") AND ("Fungi" OR "Fungal") AND ("Degradation" OR "Remediation" OR "Bioremediation"). These keywords were combined in various ways (individually and jointly) to capture a broader range of publications. This procedure allowed for the selection of previously published works related to fungal degradation of POPs in liquid cultures, soil microcosms, and field-scale applications. Additionally, methods and approaches employed to identify fungal strains with potential for bioremediation were explored. Moreover, the reference sections of retrieved articles were also examined and screened for further relevant publications. Special emphasis was placed on works published from 2003 to 2023. Exceptions were made for foundational or pioneer studies or when recent research was scarce for a specific POP. Following the search step, a thorough analysis of the data was conducted. The studies were then classified and categorized according to the POPs investigated, the fungal strains employed, and the context and conditions of the research (e.g., liquid cultures, soil microcosms, field studies). This comprehensive organization ensured all relevant aspects were addressed within the scope of this article.

## 3. Fungal bioremediation of agricultural persistent organic pollutants: close-up on species and capabilities

### 3.1. Chlorinated cyclodienes

Chlorinated cyclodienes (CDDs) are a diverse group of synthetic chlorinated hydrocarbon chemicals characterized by a polycyclic structure with variable numbers of chlorine atoms. Members of this group include among others aldrin, dieldrin, heptachlor, endrin, chlordane, toxaphene, and endosulfan. Chlorinated cyclodienes have been extensively employed for their pesticidal activities, allowing insect control on agricultural crops, as well as wood preservation as fungicides (Xiao et al. 2011). Despite the implemented restrictions and prohibition of their use between the 1970s and 1990s in many developed countries including the United States and the European Union, these pesticides remain until now present in many soil environments as a result of their low aqueous solubility, their high lipophilicity, and their structural stability, thus becoming an intrinsic part of geochemical and biological cycles (Xiao et al., 2011; Fujii et al., 2012). Thermal desorption, once considered as the most effective option for removal of these pesticides is high-cost and results in the loss of fertility in treated soils (Dicks and Goh, 2012). Microbial biodegradation of chlorinated cyclodienes using isolated bacterial and fungal strains has been explored since the 1960s. Mycoremediation has emerged as a prospective tool for the degradation of pesticides belonging to this group, offering advantages in contexts where bacterial remediation shows limitations.

The following section explores the existing knowledge on fungal degradation of the aforementioned chlorinated cyclodienes
(CDDs). We briefly summarize the mechanisms employed by fungi for the degradation of these pollutants, including the enzymatic pathways involved. We also shed light on considerations for selecting fungi for bioremediation applications for selected pollutants. Finally, we provide an overview of known fungal species reported as capable of degrading CDDs, along with their degradation percentages, thus offering a concise yet informative picture of the current state of knowledge of fungal CDD degradation research (Fig. 1).

### 3.1.1. Aldrin and dieldrin

To date, several fungal species belonging to the Ascomycota, Basidiomycota, and Mucoromycota divisions have been reported as capable of transforming aldrin and dieldrin (Fig. 1, A), namely Phlebia acanthocystis, Phlebia aurea, Phlebia brevispora (Xiao et al., 2011), Pleurotus ostreatus (Purnomo et al., 2017), Mucor racemosus strain DDF (Kataoka et al., 2010), Cordyceps brongniartii ATCC66779, Cordyceps militaris KS-92 (Xiao and Kondo, 2013), and Penicillium miczynskii CBMAI 930 (Birolli et al., 2015). While the evidence for the microbial conversion of aldrin to dieldrin via an epoxidation mechanism has been provided long ago by Lichtenstein and Schulz (Lichtenstein and Schulz, 1960), the formation of an epoxide ring structure during this process hinders further degradation of dieldrin, thus resulting in its exacerbating persistence. Therefore, fungal strains capable of transforming dieldrin are of great interest.

In these studies, fungal biotransformation of aldrin and dieldrin occurred cometabolically with glucose or dextrose as carbon sources. No study has reported to date the capacity of a fungal species to degrade aldrin and dieldrin as the sole carbon source. The proposed metabolic pathways for the transformation of aldrin and dieldrin by the fungal species displayed in Fig. 1-A are summarized in Supplementary Figure 2.

### 3.1.2. Heptachlor

While the first evidence for heptachlor degradation by soil microorganisms was reported by Miles et al. (1969), whose studies demonstrated the ability of soil fungi and bacteria to metabolize heptachlor with the concomitant production of various metabolic products such as heptachlor epoxide, chlordene, 1-hydroxychlordene, chlordene epoxide, and 1-hydroxy-2,3-epoxychlordene through

<figure><img src="image_3.png" /><figcaption>**Fig. 1.** Primary fungal species capable of degrading chlorinated cyclodienes. **Fig. 1** | References | **Aldrin and dieldrin:** Phlebia acanthocystis, Phlebia aurea, Phlebia brevispora (Xiao et al., 2011 (Xiao et al., 2011a)), Pleurotus ostreatus (Purnomo et al., 2017 (Purnomo et al., 2017)), Mucor racemosus strain DDF (Kataoka et al., 2010 (Kataoka et al., 2010), Cordyceps brongniartii ATCC66779, Cordyceps militaris KS-92 (Xiao et al., 2013 (Xiao and Kondo, 2013)), Penicillium miczynskii CBMAI 930 (Birolli et al., 2015 (Birolli et al., 2015)). | **Heptachlor:** Phlebia tremellosa, Phlebia brevispora, Phlebia acanthocystis, Phlebia aurea (Xiao et al., 2011 (Xiao et al., 2011c)), Pleurotus ostreatus*(Purnomo et al., 2013 (Purnomo et al., 2013)), Mucor racemosus strain DDF (Kataoka et al., 2010 (Kataoka et al., 2010)), Cordyceps brongniartii ATCC66779 (Xiao et al., 2013 (Xiao and Song, 2013)). | **Chlordane:** Phlebia lindtneri, Phlebia brevispora and Phlebia aurea (Xiao et al., 2011). | **Endrin:** Pleurotus pulmonarius, Pleurotus ostreatus (Sadiq, 2015; Sadiq et al., 2019a), Phlebia acanthocystis, Phlebia brevispora (Xiao et al., 2019 (Xiao and Kondo, 2019)). | **Endosulfan:** Fusarium ventricosum (Siddique et al., 2003), Trametes hirsute (Kamei et al., 2011), Aspergillus niger strain ARIFCC 1053 (Bhalerao et al., 2013), Bjerkandera adusta (Rivero et al., 2012), Paecilomyces variotii, Paecilomyces lilacinus (Hernández-Ramos et al., 2019), Aspergillus niger, Aspergillus flavus, Penicillium chrysogenum (Ahmad, 2020), consortium of Botryosphaeria laricina, Aspergillus tamarii, and Lasiodiplodia sp.(Abraham and Silambarasan, 2014), Aspergillus sydoni (Goswami et al., 2009), Mortierella sp. strain W8 (Kataoka et al., 2011; Sadiq et al., 2019b), Co-cultivation of Pleurotus eryngii and Coprinus comatus (Wang et al., 2017), Phenol hydroxylase 1PN0 of Trichosporon cutaneum (Singh et al., 2019).</figcaption></figure>
various metabolic pathways, a significant dearth of research exists regarding the ability of fungal strains to transform this compound. Nonetheless, a selection of ligninolytic and non-ligninolytic fungal strains have been demonstrated to degrade heptachlor and its more recalcitrant degradation product heptachlor epoxide (Fig. 1, B), (Supplementary Figure 3). These fungal strains include: *Phlebia tremellosa*, *Phlebia brevispora*, *Phlebia acanthocystis*, *Phlebia aurea* (Xiao, Mori, Kamei, and Kondo 2011a), *Pleurotus ostreatus* (Purnomo et al., 2013), *Mucor racemosus* strain DDF (Kataoka et al., 2010), and *Cordyceps brongniartii* ATCC66779 (Xiao and Song, 2013).

### 3.1.3. Chlordane

Despite ranking 22nd in the ATSDR's Priority List of Hazardous Substances and figuring among the 12 initial POPs by the Stockholm Convention, very limited information is available on the fungal biotransformation of the organochlorine pesticide chlordane. Nevertheless, a study conducted by Xiao et al. (2011) reported that three *Phlebia* species, *P. lindtneri*, *P. brevispora*, and *P. aurea* were capable of metabolizing *trans*-chlordane, the major constituent of technical chlordane. *Phlebia lindtneri* demonstrated the highest ability to transform *trans*-chlordane, with 47 % of dieldrin being degraded after a 14-day incubation period, and about 64 % after 42 days, whereas *P. brevispora*, and *P. aurea* degraded about 50 % of dieldrin following 42 days of incubation. Degradation of *trans*-chlordane by these three *Phlebia* species was accompanied by the accumulation of products such as oxychlordane, heptachlor, heptachlor epoxide, dichlorochlordene, chlordane chlorohydrin, monohydroxylated chlordene, heptachlor diol, dihydroxychlordene, and three unidentified products, with heptachlor epoxide being the major product of *trans*-chlordane transformation. In this same study, lower degradation activity along with a reduced number of metabolic products was reported in cultures confronted with high concentrations of P450 inhibitors, thus demonstrating the involvement of the intracellular cytochrome P450 monooxygenases in some routes of *trans*-chlordane metabolism, which suggests that hydroxylation reactions are of major importance in *trans*-chlordane biotransformation. Additional fungal species that have demonstrated the ability of transforming chlordane are resumed in Fig. 1, C.

### 3.1.4. Endosulfan

Endosulfan is a broad-spectrum organochlorine pesticide characterized by the presence of a sulfur moiety. In its commercial form, endosulfan is a mixture of two stereoisomers, α- and β-endosulfan, combined at a ratio of 7:3. This pesticide has been extensively and widely used for over 30 years on a variety of crops such as rice, cotton, fruits, vines, vegetables, tea plantation, and ornamental plants, which resulted in its widespread accumulation in soil environments (Abraham and Silambarasan, 2014).

Contrarily to what has been seen in the case of other chlorinated cyclodiene pesticides, prohibition and banning of endosulfan didn't occur until recently following its inclusion in the Stockholm Convention list of POPs in 2011. Nonetheless, many countries such as India and China continue to use endosulfan to date, therefore its persistence in various environments and its possible bio-accumulation into the food chain are still of great concern. At high concentrations, endosulfan has been shown to have detrimental effects on soil fungal communities, affecting both their growth and species biodiversity, thus shifting microbial communities towards tolerant species. Conversely, at low concentrations, endosulfan has been shown to exert a stimulating effect on microbial growth, acting possibly as a sulfur and carbon source (Bisht et al., 2014).

Degradation of endosulfan by fungi species has been reported in relation to the genera *Aspergillus* (Abraham and Silambarasan, 2014; Ahmad, 2020; Barberis et al., 2019; Bhalerao, 2013; Bhalerao and Puranik, 2007; Goswami et al., 2009; Hussain et al., 2006; Mukherjee and Mittal, 2005; Hernández-Ramos et al., 2019), *Penicillium* (Guillén-Jiménez et al., 2012; Quintero et al., 2008), *Fusarium* (Siddique et al., 2003), *Trametes* (Bisht et al., 2019), *Cladosporium* (Nagpal et al., 2008; Quintero et al., 2008), *Trichosporon* (Singh et al., 2019), *Pleurotus* (Sadiq et al., 2019a; Sadiq et al., 2019b; Wang et al., 2017), *Paecilomyces* (Hernandez-Ramos et al., 2019), *Botryosphaeria* (Abraham and Silambarasan, 2014), *Lasiodiplodia* (Abraham and Silambarasan, 2014), *Bjerkandera* (Rivero et al., 2012), *Mortieralla* (Kataoka et al., 2011), *Chaetosartorya* (Hussain et al., 2006), and *Coprinus* (Wang et al., 2017) (Fig. 1, E), (Supplementary Figure 4).

Most studies addressing fungal degradation of endosulfan have reported a significant decrease in the pH of culture media (Ahmad, 2020; Bhalerao, 2013; Bhalerao and Tejomyee, 2012; Siddique et al., 2003; Hernández-Ramos et al., 2019). This acidification can be retraced to the release of chloride ions resulting from the dehalogenation of endosulfan, leading to the possible formation of HCl in the culture medium. In addition, the degradation of endosulfan by some fungal species results in the release of organic acids as metabolic products, which can further acidify the medium (Siddique et al., 2003). Nonetheless, some other studies reported no significant pH variation following the fungal transformation of endosulfan (Sadiq et al., 2019a; Kataoka et al., 2011; Abraham and Silambarasan, 2014).

### 3.1.5. Endrin

Endrin is another organochlorine cyclodiene insecticide that was vastly employed for the control of pests, notably on cotton and rice sugar cane, before its prohibition in the 1970s as a result of its toxicity, teratogenicity, and its extreme persistence in soil environments and sediments (Xiao and Kondo, 2019). Nevertheless, its prominent detection in agricultural zones around the world long after its banning (Gopalan and Chenicherry, 2018), encourages the continuous search for appropriate eco-friendly remediation technologies that could replace physicochemical methods regarded as intrusive for affected soils and agricultural zones. Studies addressing fungal transformation of endrin are presented in Fig. 1, D, and mainly include some white-rot fungi, namely *Pleurotus pulmonarius*, *Pleurotus ostreatus* (Sadiq, 2015), *Phlebia acanthocystis*, *Phlebia brevispora* (Xiao and Kondo, 2019).
## 3.2. Other non-cyclodiene insecticides

### 3.2.1. Dichlorodiphenyltrichloroethane (DDT)

Despite its banning for agricultural purposes, DDT and its metabolic products continue to be detected in several environmental matrices including soil (Malusá et al. 2020), and honey (Cervera-Chiner et al., 2020), consequently impacting human health through the food chain. The detection of DDT in several matrices long after its banning can be traced to its continuous use in tropical countries as part of public health programs for malaria control, in addition to its intrinsic capacity to travel long distances and persist long after its introduction.

During the past decades, bioremediation of DDT has been addressed to a large extent as part of the attempts aiming to elaborate cost-effective, environmentally-friendly remediation technologies. Among others, mycoremediation has shown to be of increasing interest as a result of the ability of fungal species to withstand harsh environmental conditions and degrade a large array of xenobiotic compounds with regard to their extracellular enzymatic systems. The ability to degrade DDT and its metabolic products has been reported in numerous fungal species belonging to the genera Phanerochaete (Bumpus and Aust, 1987), Fusarium (Mitra et al., 2001), Gloeophyllum, Fomitopsis (Cea et al., 2010; Souza et al., 2011), Daedalea (Cea et al., 2010; Souza et al., 2011; Rubilar et al., 2011), Phlebia (Xiao et al., 2011b), Trametes (Sari et al., 2012), Flammulina (Fan et al., 2013), Nectria (Thomas and Gohil, 2010), Trichoderma (Hechmi et al., 2016; Carvalho et al., 2011), Ganoderma (Rubilar et al., 2011; Carvalho et al., 2013). More detailed information on the degradation of DDT is provided in Table 1 (section color code: blue), where we have listed the fungal species involved or shown potential in its biodegradation. This table also displays the percentage of degradation, the incubation conditions (time and temperature), the culture medium or environment used, the degradation products identified, as well as additional notes. Additionally, fungal degradation pathways of this pollutant are summarized in Supplementary Figure 5.

### 3.2.2. Hexachlorobenzene

Hexachlorobenzene (HCB) is an organochlorine pesticide that was widely used as a fungicide for wood preservation and control of common bunt on wheat crops. Its environmental toxicity and risk of bioaccumulation and biomagnification in the food chain resulted in its enlistment in the Stockholm convention list of POPs. Despite its banning since 1965, its occurrence to this date can be attributed to its formation as a byproduct during the manufacture of chemical solvents and pesticides. Little information is available concerning fungal remediation of hexachlorobenzene. Fungal species such as *Psilocybe cf. castanella* strain CCB444, *Trametes villosa* strain CCB176, and *Eupenicillium crustaceum* SSP1953 have been reported to exhibit varying degrees of tolerance and degradation potential towards HCB. A more detailed presentation of the aspects related to the biodegradation of HCB is available in Table 1 (section color code: yellow), where we have listed the fungal species involved and their performance in terms of percentage of degradation. This table also specifies the incubation conditions (time and temperature), the culture medium or environment used, the degradation products, as well as additional remarks regarding fungal degradation of this pollutant.

### 3.2.3. Chlordecone

The chlorinated cyclodiene chlordecone (1,1a,3,3a,4,5,5,5a,5b,6-decachlorooctahydro-1,3,4-metheno-2H-cyclobuta[cd]pentalen-2-one), is an insecticide that was largely used in the French West Indies to control the banana root borer, *Cosmopolites sordidus*, for over an 18-year period (Merlin et al., 2014). Structurally, chlordecone is a bis-homocubane compound consisting of a cage structure comprising 10 chlorine atoms in addition to a ketone functional group. Decades following its prohibition in mainland France and the Antilles, notable chlordecone concentrations were detected in soils previously used for banana plantations, as well as in surrounding aquatic environments and food crop products, thus evincing its environmental persistence and biomagnification potential ("Impacts de l'utilisation de La Chlordécone et Des Pesticides Aux Antilles: Bilan et Perspectives d'évolution - Sénat (n.d.)"). In addition, exposure to chlordecone has been associated with neurological impairment and an increased risk of cancer in affected populations (Merlin et al., 2014). Modeling studies have predicted the persistence of this insecticide for about 60–100 years in nitisols, 3–4 centuries in ferralsols, and up to 7 centuries in andosols of the Antilles heavily impacted by chlordecone contamination between 1972 and 1993. It is therefore imperative to intensify the efforts aiming to develop new, more environment-friendly depollution methods, allowing to reduce the burden of chlordecone contamination and its repercussions on humans, animals, and aquatic life via the food chain. Nevertheless, due to its precocious banning in the United States, which preceded its extensive use in Guadeloupe and Martinique, and therefore limited its pollution to the French Antilles, a few numbers of studies have addressed the potential ability of microorganisms to biotransform chlordecone and even fewer have addressed fungal bioremediation potential.

Merlin et al. (2014) reported the ability of several fungal genera belonging to the phylum Ascomycota to tolerate and grow in the presence of chlordecone. However, *Fusarium oxysporum* strain MIAE01197 emerged as the only strain able to tolerate and degrade up to 40 % of chlordecone in liquid culture medium, where chlordecone served as the sole carbon source. In addition, chlordecone was shown to exert a stimulatory effect on *F. oxysporum* MIAE01197 growth in the presence of glucose as the carbon source. No metabolic products resulting from chlordecone transformation were detected in this study. This study represents the only work addressing fungal remediation of chlordecone, whereas other available studies tackling chlordecone bioremediation investigated the impact of chlordecone on soil microbial diversity (Merlin et al., 2016), or evaluated anaerobic microbial transformation as a potential method for chlordecone transformation (Lomheim et al., 2020). Previous studies have suggested that some bacterial species, such as *Methanosarcina thermophila* and strains of *Pseudomonas* spp., possess the ability to degrade chlordecone. For example, Jablonski et al. (1996) reported that *M. thermophila* could convert chlordecone to various polar and nonpolar metabolites under methanogenic conditions, with 86 % dechlorination of the initial [¹⁴C]-chlordecone. The authors suggested that cobalamins and F430, present in all methanogens, could play a key role in this transformation, implying that all methanogens could potentially attack chlordecone. Furthermore,
<table><caption>**Table 1**
Primary fungal species capable of degrading the four agricultural POPs (color coding DDT (in blue), Hexachlorobenzene (in yellow), Lindane (in grey) and Pentachlorophenol (in pale peach)).</caption><thead><tr><th>Fungal Species</th><th>Degradation %</th><th>Incubation</th><th>Medium</th><th>Degradation Products</th><th>Notes</th><th>Reference</th></tr></thead><tbody><tr><td>Phanerochaete chrysosporium BKM-F-1767</td><td>74% ±67%</td><td>18d, 39°C</td><td>Nitrogen-Limited Liq Medium</td><td>DDE, DDD, DBP, Dicofol, FW-152</td><td>Mass Balance Analysis, Capable of degrading DDE, Involvement of Ligninolytic Systems (LMEs)</td><td>(Bumpus et al., 1993)</td></tr><tr><td rowspan="3">Pleurotus ostreatus</td><td>48%</td><td>28d, 25°C</td><td rowspan="3">Spent Mushroom Waste<br/>Sterilized Soils (SL)<br/>Unsterilized Soils (USL)</td><td rowspan="3">NM*</td><td rowspan="3">LMEs not involved<br/>Mineralization 5.1%<br/>Mineralization 8%</td><td rowspan="3">(Purnomo et al., 2010)</td></tr><tr><td>40%</td><td>28d, 25°C</td></tr><tr><td>80%</td><td>28d, 25°C</td></tr><tr><td>Pleurotus ostreatus OST-1</td><td>67%</td><td>25°C</td><td>Liquid Malt Extract</td><td>NM</td><td></td><td>(Sadiq, 2015)</td></tr><tr><td>Nectria mariannaeae FJ940904</td><td>50%</td><td>3 weeks</td><td>Microcosm</td><td></td><td>Capable of degrading DDE and DDD</td><td>(Thomas and Gohil, 2010)</td></tr><tr><td>Phlebia lindneri GB-1027</td><td>70%</td><td></td><td rowspan="2">Low Nitrogen Medium<br/>+ Glucose</td><td rowspan="2">DDD, DDA, DBP,<br/>and DBH</td><td rowspan="2">DDE not detected,<br/>Involvement of P450 Monooxygenase,<br/>Degradation through hydroxylations and ring cleavage</td><td rowspan="2">(Xiao et al., 2011b)</td></tr><tr><td>Phlebia brevispora TMIC34596</td><td>30%</td><td>21d, 30°C</td></tr><tr><td>Trametes versicolor U97</td><td>73%</td><td>40d</td><td>Malt Extract Medium</td><td>DDE, DDD, DDMU,<br/>4,4'-DCBP and 4-CBA</td><td>Co-Metabolism,<br/>LMEs involved (Li-P)</td><td>(Sari et al., 2012b)</td></tr><tr><td>Gloeophyllum trabeum NBRC6509</td><td>87%</td><td></td><td></td><td>DDE, DDD, and DBP</td><td>Co-Metabolism,<br/>Involvement of Fenton-type reactions,<br/>Increase in the pH</td><td>(Purnomo et al., 2008)</td></tr><tr><td rowspan="2">Daealea dickinsii NBRC31163</td><td rowspan="2">81%</td><td rowspan="2">28d, 30°C</td><td rowspan="2">Potato Dextrose Broth (PDB)</td><td rowspan="2">NM</td><td rowspan="2">None</td><td rowspan="2"></td></tr><tr><td></td></tr><tr><td>Fomitopsis pinicola NBRC8705</td><td>84%</td><td></td><td></td><td>NM</td><td>None</td><td></td></tr><tr><td>Gloeophyllum trabeum NBRC6509</td><td>41% (SL)<br/>43% (USL),<br/>64% (HCS)</td><td></td><td rowspan="3">Artificially contaminated<br/>sterilized soils (SL),<br/>unsterilized soils (USL), and<br/>historically contaminated soil<br/>(0.8 mmol g-1) (HCS)</td><td rowspan="3">DDD (in USL)</td><td rowspan="3">Gloeophyllum trabeum can be applied directly for the<br/>degradation of DDT in soil without supplementary<br/>treatments</td><td rowspan="3">(Purnomo et al., 2011)</td></tr><tr><td>Daealea dickinsii NBRC31163</td><td>15% (SL),<br/>32% (USL)</td><td rowspan="2">28d, 30°C</td></tr><tr><td>Fomitopsis pinicola NBRC8705</td><td>9% (SL),<br/>32% (USL)</td></tr><tr><td>Flammulina velutipes + Laccase</td><td>67%</td><td>21d, 28°C,<br/>pH=4.5</td><td>Microcosm Soils (natural and<br/>sterilized soils)</td><td>DDOH, DDMS</td><td>T=28°C, and pH=4.5 are optimal for co-remediation,<br/>No significant difference in DDT degradation between<br/>sterilized and natural soils,<br/>Increasing concentrations of Cd2+ exerted an<br/>inhibitory effect on DDT remediation,<br/>6Ug=1 soil of laccase is the optimal concentration</td><td>(Fan et al., 2013b)</td></tr><tr><td>Pleurotus eryngii NBRC 32798 + Ralstonia picketti<br/>+ Pseudomonas aeruginosa</td><td>43% (pure culture),<br/>78% (co-culture<br/>with R. picketti)<br/>82% (co-culture<br/>with P. aeruginosa)</td><td>7d, 30°C</td><td>Potato Dextrose Broth (PDB)</td><td>DDE, DDD, DDMU<br/>(with R. picketti)<br/>DDE and DDD (with P.<br/>aeruginosa)</td><td>Synergistic activity with<br/>Ralstonia picketti and Pseudomonas aeruginosa,<br/>Biosurfactant-producing bacteria increase the solubility<br/>of DDT</td><td>(Purnomo et al.,<br/>2019)<br/>(Maulianawati et<br/>al., 2021)</td></tr><tr><td>Daealea dickinsii NBRC 31163<br/>+ Ralstonia picketti</td><td>53.6% (pure<br/>culture),<br/>68.6% (co-culture)</td><td>7d, 30°C</td><td>Potato Dextrose Broth (PDB)</td><td>DDD and DDE (pure<br/>cultures of D. dickinsii),<br/>DDD and DDMU<br/>(co-cultures)</td><td>Ralstonia picketti enhance the degradation of DDT by<br/>Daealea dickinsii</td><td>(Setyo et al., n.d.)</td></tr></tbody></table>

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<table><caption>Table 1 (continued)</caption><thead><tr><th></th><th></th><th></th><th></th><th></th><th></th><th></th></tr></thead><tbody><tr><td>Ganoderma lingzhi BMC9057</td><td>53% (pure culture), 82% (co-culture with B. subtilis) 100% (co-culture with P. aeruginosa)</td><td>7d, 30°C</td><td>Potato Dextrose Broth (PDB)</td><td>DDD and DDMU</td><td>Bacillus subtilis and Pseudomonas aeruginosa enhance the degradation of DDT by Ganoderma lingzhi</td><td>(Grizca Boelan et al., 2018) (Boelan and Purnomo, 2019)</td></tr><tr><td>Trichoderma hamatum FBL 587</td><td>84% (co-culture with Cucurbita pepo)</td><td>≈ 40d, 25°C</td><td>Pot Experiment (Unsterilized Soil)</td><td>DDE</td><td>Upregulation of genes encoding Phase I, Phase II, and Phase III enzymes such as cytochrome P450 monooxygenases, epoxide hydrolases, glutathione S-transferases, glyoxyltransferases, sulfotransferases, in addition to transmembrane transporters including ABC and MFS transporters</td><td>(Davolos et al., 2021)</td></tr><tr><td>Pseudocne cf. castanella CCB444</td><td>10.80%</td><td></td><td></td><td></td><td></td><td></td></tr><tr><td>Lentinus cf. zeyheri CCB274</td><td>4.80%</td><td>65d, 30°C</td><td>Microcosm Soils (Unsterilized)</td><td>NM</td><td>None</td><td>(Matheus et al., 2000)</td></tr><tr><td>Eupenicillium crustaceum SSP1953</td><td>25%</td><td>56d, 28 ± 2°C</td><td>Microcosm Soils (Sterilized) + sugar cane bagasse</td><td>Non-extractable ¹⁴C-residues were formed</td><td>None</td><td>(Vitali et al., 2006)</td></tr><tr><td>Pleurotus eryngii</td><td>49% (MG), 17.5% (SL)</td><td>10d, 30°C</td><td>Liquid Culture Medium (MG), Soil Microcosms (Sterilized SL and Unsterilized USL)</td><td>No detectable metabolites, complete mineralization ?</td><td>A panel of WRF were screened for their ability to degrade isomers of HCH, α-HCH isomer was easily degraded, β-HCH isomer resisted the most to degradation, δ-HCH exhibited the most detrimental effects on fungal growth</td><td>(Quintero et al., 2008)</td></tr><tr><td>Bjerkandera adusta</td><td>42% (MG)</td><td></td><td></td><td>3-chloro-4-methoxyphenyl ethanone, 2,4-dichloro-3-metoxibenzene-carbonyl chloride</td><td>α-, β-, δ-, and γ-HCH were investigated for their degradation by P. florida,</td><td></td></tr><tr><td>Pleurotus florida</td><td>100% (Raper's medium), 61-67% (Soil Microcosm, 60d)</td><td>30d, 25°C</td><td>Raper's complete medium, Soil Microcosm + SMS</td><td>No detectable metabolites, complete mineralization ?</td><td>P. florida degraded all HCH isomers equally efficiently in medium, γ-HCH degradation half-life was shortened from 439-570 days in un-amended soil to 37-42 days in P. florida SMS-amended soil</td><td>(Mohapatra et al., 2015)</td></tr><tr><td>Trametes hirsutus</td><td>≈ 95%</td><td></td><td></td><td></td><td>At a concentration of 0.27 µmol l⁻¹, Lindane exhibits no detrimental effects on fungal growth, No involvement of intracellular enzymes during the first degradation steps, Same metabolic pathway employed in both fungi</td><td></td></tr><tr><td>Phanerochaete chrysosporium</td><td>≈ 90%</td><td>28d</td><td>Brodies Liquid Medium (+VA)</td><td>TCCH and TCCOL</td><td></td><td>(Singh and Kuhad, 1999)</td></tr><tr><td>Phanerochaete sordida HHB8922-50</td><td>≈ 90%</td><td>28d, 28°C</td><td>Brodies Liquid Medium (+VA)</td><td>TCCH and TCCOL</td><td></td><td></td></tr><tr><td>Cyathus bulleri 195062</td><td>≈ 95%</td><td></td><td>Peptone Dextrose Liquid Medium (+VA)</td><td>TCCOL</td><td>At a concentration of 0.27 µmol l⁻¹, Lindane exhibits no detrimental effects on fungal growth, No involvement of intracellular enzymes during the first degradation steps</td><td>Singh and Kuhad, 2000</td></tr><tr><td>Phanerochaete chrysosporium BKM-F-1767</td><td>≈ 50%</td><td>12d, 25-37°C</td><td>Liquid growth medium</td><td>TCCH, TCCOL, and TCCE</td><td>No involvement of peroxidases in lindane transformation, Possible involvement of cytochrome P450 monooxygenase</td><td>Mougin et al., 1996</td></tr><tr><td>Pleurotus ostreatus</td><td>25.8 mg/g/g (degraded lindane/ initial lindane/final biomass)</td><td>12.45d, 24.5°C, N concentration of 1.89 g/L</td><td>Basal Liquid Mineral Medium</td><td>NM</td><td>P. ostreatus produced, Laccase, Mg-P, but not Li-P as part of its ligninolytic arsenal</td><td>(Rigas et al., 2005)</td></tr></tbody></table>

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<table><caption>Table 1 (continued)</caption><thead><tr><th></th><th></th><th></th><th></th><th></th><th></th></tr></thead><tbody><tr><td>Trametes versicolor ZIM L</td><td>98%</td><td></td><td></td><td>No chlorinated products were detected</td><td>No substantial lindane degradation by the WRF Chondrosierum purpureum, The BRF Gloeophyllum trabeum was not able to degrade lindane, Lindane degradation mostly occurred between 6h and 1d of exposure</td></tr><tr><td>Hypoxylon fragtiforme ZIM L108</td><td>&gt; 90%</td><td>21d, 25°C</td><td>Liquid Culture Medium</td><td></td><td></td></tr><tr><td>Pleurotus ostreatus ZIM L</td><td>&gt; 90%</td><td></td><td></td><td></td><td></td></tr><tr><td>Phlebia brevispora TMIC34596</td><td>87% (LN)</td><td></td><td>Low-nitrogen (LN) and High-nitrogen (HN) Liquid Media</td><td>TCCH, TCCOL, PCCOL, TCCDL, THTCH, Tri-CCDL</td><td>Involvement of cytochrome P450 monooxygenase in the degradation, Mineralization of lindane into carbon dioxide was observed, Mineralization into CO2 increased with addition of glucose</td></tr><tr><td>Phlebia lindneri GB1027</td><td>73% (LN)</td><td>25d, 30°C</td><td></td><td></td><td></td></tr><tr><td>Conidiobolus 03-1-56</td><td>100%</td><td>25d, 28°C</td><td>MGYP Broth</td><td>Complete absence of degradation metabolites</td><td>Conidiobolus 03-1-56 was unable to utilize and degrade lindane as the sole carbon and energy source (co-metabolism), Carbon- and nitrogen-limiting conditions, + addition of VA enhanced Lindane degradation, Involvement of LMEs in the degradation, Adsorption was not involved in Lindane removal</td></tr><tr><td>Fusarium solani</td><td>59.40%</td><td>10d, 28 ± 2°C</td><td>C-ZapekDox Liquid Medium</td><td>NM</td><td>Maximum growth inhibition was observed at a concentration of 600 µg ml<sup>-1</sup></td></tr><tr><td>Fusarium poae</td><td>56.70%</td><td></td><td></td><td></td><td></td></tr><tr><td>Fusarium verticillioides AT-100</td><td>81%</td><td>7d, 30 ± 2°C</td><td>Minimal Salt Medium (MSM)</td><td>γ-2,3,4,5,6-PCCH, γ-1,3,4,5,6-PCCH, Benzoic acid, and Hydroxybenzoic acid</td><td>Strain isolated from leaves of Agave tequilana, Lindane was used as the sole carbon and energy source, Degradation was enhanced under nitrogen- and phosphorus-limited conditions, in the presence of A. tequilana leaves, and with addition of copper and yeast extract</td></tr><tr><td>Rhodotorula sp. VITJzN03</td><td>100%</td><td>10d, 30°C, pH=6</td><td>Sterile Mineral Medium</td><td>γ-PCCH, 1,4-TCHDiene, 1,4 DCB, 1,2,4 TCB, CDCHdiene and MA</td><td>Lindane was used as the sole carbon and energy source, Complete mineralization occurred at a concentration of 600 mg/L</td></tr><tr><td>Candida VITJzN04</td><td>100%</td><td>8d, 30°C, pH=7</td><td>Mineral Medium (MM)</td><td>γ-PCCH, TCCH-1-ol, 2,5-DCHQ, 2-CHQ, HMSA, and MA</td><td>Lindane was used as the sole carbon and energy source, Complete degradation occurred at a concentration of 600 mg/L, H2O2 addition significantly increased degradation rates, Involvement of dichlorohydroquinone reductive dechlorinases, dechlorinases, dehalogenases, Li-P and Mn-P was observed</td></tr><tr><td>Phanerochaete chrysosporium BKM-F-1767</td><td>100% (100 mg/L)</td><td>18d, 39°C</td><td>Microcosm Soil (Non-Sterile)</td><td>PCA</td><td>Among the first PCP-degrading fungal species reported, Mineralization rates revealed a linear increase with increasing concentrations of the pesticide (up to 1600 mg/L).</td></tr><tr><td>Lentinula edodes LE2</td><td>99% (SL)</td><td>10 weeks, 25°C</td><td>Microcosm Soil (Non-Sterile NSL, Sterile SL)</td><td>PCA, TeCA, TeCP</td><td>Efficient as monocultures, Degradation was accompanied by a pH decrease. In non-sterile soils as mixed cultures, degradation of pentachlorophenol was markedly reduced and was accompanied by a significantly reduced activity of phenol oxidase and Mn-P</td></tr></tbody></table>

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<table><caption>Table 1 (continued)</caption><thead><tr><th></th><th></th><th></th><th></th><th></th><th></th><th></th></tr></thead><tbody><tr><td>Trametes versicolor PRL 572</td><td>96% (SL)</td><td>42d, 28°C</td><td>Microcosm Soil (Non-Sterile) + Straw</td><td>PCA, TeCA</td><td>29% of PCP was mineralized, Only trace amounts of PCA and TeCA were detected, Possible involvement of laccases?</td><td>(Tuomela et al., 1998)</td></tr><tr><td>Coriolus versicolor FPRL-28A (Laccase)</td><td>100% (25 µg ml<sup>-1</sup>), 60% (200 µg ml<sup>-1</sup>)</td><td>72h, 25°C, pH=5</td><td>100 mM sodium acetate buffer, replacing catechol by PCP</td><td>o- and p-chloranilis (tetrachlorobenzoquinones)</td><td>None</td><td>(Ullah et al., 2000)</td></tr><tr><td>Deconica castanella CCIIB278 (Laccase)</td><td>23% (100 µg L<sup>-1</sup>)</td><td>24h, 25°C, pH=4.5</td><td>50 mM sodium acetate buffer</td><td>NM</td><td>D. castanella Lac is a thermostable enzyme</td><td>(Coelho et al., 2019)</td></tr><tr><td>Pleurotus pulmonarius CCB19</td><td>70%</td><td>5d, 28°C</td><td>Mineral Solution + corn cob extract (Lac inducer)</td><td>NM</td><td>Laccase played a primordial role in the degradation of PCP</td><td>(de Souza et al., 2011)</td></tr><tr><td rowspan="2">Trametes versicolor HR131</td><td>1300-1800 mg of PCP/kg of soil</td><td>42d, 23 ± 3°C</td><td>Soil Microcosm (Aged -contaminated > 30 years)</td><td>NM</td><td>Aged PCP contamination in field-contaminated soils turns to be more toxic to WRF than laboratory spiked PCP soils</td><td>(Walter et al., 2005b)</td></tr><tr><td>From 800-100 mg kg<sup>-1</sup> to < 50 mg kg<sup>-1</sup></td><td>74 weeks, T<sub>0</sub> x=35°C</td><td>Field-scale experiments (Aged -contaminated soil)</td><td>NM</td><td>T. versicolor can degrade PCP in aged-contaminated soils on a field scale, T. versicolor was introduced in soil via engineered soil cells</td><td>(Walter et al., 2005a)</td></tr><tr><td rowspan="2">Trametes hirsutu</td><td>70% (1043 mg/kg)</td><td>7 weeks, 23°C</td><td>Soil Microcosm (Aged -contaminated soil)</td><td>PCA</td><td>Poplar sawdust more suitable substrate than fir and pine sawdust, Inoculum age, presence/absence of a co-substrate and growth substrate composition affect PCP removal</td><td>(Ford et al., 2007)</td></tr><tr><td>3.8% day<sup>-1</sup> (Liquid) 1.6% day<sup>-1</sup> (Sawdust)</td><td>8d, 22°C</td><td>N Liquid Medium and Sawdust culture</td><td>NM</td><td>PCP-mineralization and growth remained high between pH 3-7.5, Mineralization increased by increasing pH value above the pKa, The phenol form of PCP, which forms at lower pH values, is more toxic on fungal growth of T. hirsutu compared to the phenate form of PCP</td><td>(Boyle, 2006)</td></tr><tr><td>Anthracophyllum discolor</td><td>96% (100 mg PCP/Kg of soil) 94% (250 mg PCP/Kg of soil), 81% (250 mg PCP/Kg of soil in soil Slurry Reactor)</td><td>4 weeks, 30°C, pH=4.5</td><td>Soil Microcosm (Slurry soil flasks cultures + Distilled dried grains with solubles DDGS), Soil Slurry Reactor (SL)</td><td>PCA, 2,5-dichloro-1,4-dimethoxybenzene, 2-chloro-1,4-dimethoxybenzene, and 3,4-imethoxybenzaldehyde</td><td>The use of DDGS enhances PCP degradation, No involvement of the ligninolytic system by A. discolor</td><td>(Rubilar et al., 2007)</td></tr><tr><td>Bjerkandera adusta ATTC 90940</td><td>93% (100 mg PCP/Kg of soil) 70% (250 mg PCP/Kg of soil)</td><td></td><td></td><td>NM</td><td></td><td></td></tr><tr><td>Anthracophyllum discolor Sp4</td><td>86% (25 mg/L)</td><td></td><td></td><td>NM</td><td>A. discolor displayed the greatest ability to produce Li-P and Mn-P, No correlation between the ability to degrade PCP and the tolerance ability of a screened species</td><td>(Tortella et al., 2008)</td></tr><tr><td>Lenzites betulina RU-30</td><td>91% (25 mg/L)</td><td>15d, 25°C</td><td>CLN liquid medium (N-limited, under ligninolytic conditions)</td><td>NM</td><td></td><td></td></tr><tr><td>Galerina patagonica Sp3</td><td>86% (25 mg/L)</td><td></td><td></td><td>NM</td><td>Correlation between strongest bioremediation potential and tolerance, Degradation occurred through co-metabolism (glucose) for most fungal species,</td><td>(Carvalho et al., 2009)</td></tr><tr><td>Penicillium brevicompactum Dierckx</td><td>58% (10 mg/L)</td><td>60d, 27°C</td><td>Minimal medium (MM) and glucose minimal medium (GMM) in 96-well microtitre plates</td><td>DCBQ</td><td></td><td></td></tr><tr><td>Penicillium olsonii DSM 16515</td><td>52% (10 mg/L)</td><td></td><td></td><td>NM</td><td></td><td></td></tr></tbody></table>

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<table><caption>Table 1 (continued)</caption><thead><tr><th></th><th></th><th></th><th></th><th></th><th></th></tr></thead><tbody><tr><td>Penicillium janczewskii</td><td>61% (10 mg/L)</td><td></td><td>CHQ</td><td>M. plumbeus, P. glandicola, C. sitophila, P. janczewskii, and T. longibrachiatum degraded PCP as the sole carbon source</td><td></td></tr><tr><td>Penicillium glabrum</td><td>45% (10 mg/L)</td><td></td><td>CHQ</td><td></td><td></td></tr><tr><td>DSM 16516</td><td></td><td></td><td></td><td></td><td></td></tr><tr><td>Eupenicillium hirayamae</td><td>57% (10 mg/L)</td><td></td><td>NM</td><td></td><td></td></tr><tr><td>Trichoderma longibrachiatum</td><td>100% (10 mg/L)</td><td></td><td>NM</td><td></td><td></td></tr><tr><td>Mucor plumbeus</td><td>100% (10 mg/L)</td><td></td><td>NM</td><td></td><td></td></tr><tr><td>Penicillium glandicola Oudem</td><td>81% (10 mg/L)</td><td></td><td>CHQ</td><td></td><td></td></tr><tr><td>Chrysonilia sitophila</td><td>79% (5 mg/L)</td><td></td><td>CHQ</td><td></td><td></td></tr><tr><td>Anthracophyllum discolor</td><td>93% (250 mg PCP kg-1 soil)</td><td>28d, 25°C</td><td>NM</td><td>Wheat grains more suitable substrate than wheat straw and wood chips, Immobilization in wheat grains enhanced fungal colonization, Mn-P production, and PCP degradation ability in soil, In non-sterile soils, microbial competition decreased when inoculated fungi were previously immobilized in wheat grains</td><td>(Rubilar et al., 2011)</td></tr><tr><td>Phanerochaete chrysosporium</td><td>79% (250 mg PCP kg-1 soil)</td><td></td><td>NM</td><td>At high conc (100 mg/L) PCP causes collapse of fungal spores and hyphae, C. elegans SD12, SD14 and Trichoderma SD19, SD29 exhibited tolerance to high conc of PCP</td><td>(Sing et al., 2014)</td></tr><tr><td>Cunninghamella elegans</td><td>51.7% (20 mg/L)</td><td>15d, Room T°C</td><td>NM</td><td>98% removal when both fungi were used simultaneously</td><td>(Bosso et al., 2015)</td></tr><tr><td>UMAS SD12</td><td></td><td></td><td>NM</td><td>Co-inoculation of both strains → Synergistic effect removing 95% (12.5 mg kg-1) and 80% (25 mg kg-1) of PCP</td><td>(Hechmi et al., 2016)</td></tr><tr><td>Byssochlamys nivea</td><td>90% (+ compost)</td><td>28d, 25°C, 60% moisture</td><td>NM</td><td></td><td></td></tr><tr><td>Scopulariopsis brumptii</td><td>92% (+ compost)</td><td></td><td>NM</td><td></td><td></td></tr><tr><td>Byssochlamys nivea</td><td>60%</td><td>28d, 25°C</td><td>NM</td><td></td><td></td></tr><tr><td>Scopulariopsis brumptii</td><td>62%</td><td></td><td>NM</td><td></td><td></td></tr><tr><td>Trichoderma harzianum</td><td>100%</td><td>7d, 32°C</td><td>3% Malt Liquid Medium (20 mg L<sup>-1</sup> of PCP)</td><td>Capable of degrading PCA and 2,3,4,6-TeCA</td><td>(Vacondio et al., 2015)</td></tr><tr><td>CBMAI 1677</td><td></td><td></td><td>PCA and 2,3,4,6-TeCA</td><td></td><td></td></tr><tr><td>Stropharia rugosoanulata</td><td>73% (Degraded) and 26% (Mineralized)</td><td>15 weeks</td><td>Soil Microcosm (Non-sterile)</td><td>Ability to efficiently colonize soils</td><td>(Valentin et al., 2013)</td></tr><tr><td>Phlebia acanthocystis</td><td>100%</td><td>15d, 30°C</td><td>Low-nitrogen Liquid Medium (LN)</td><td>The ability of P. acanthocystis to transform PCP significantly decreased under nutrient-rich conditions, Involvement of both ligninolytic enzymes, particularly Li-P, and the intracellular CYP450 monoxygenase</td><td>(Xiao and Kondo, 2020b)</td></tr></tbody></table>

**Abbreviations.** Not mentioned (NM), Dichlorodiphenyldichloroethylene (DDE), Dichlorodiphenyldichloroethane (DDD), Dichlorobenzophenone (DBP), 2,2-dichloro-1,1-bis(4-chlorophenyl) ethanol (FW-152), 2,2-bis(4-chlorophenyl)acetic acid (DDA), 4,4-dichlorobenzhydrol (DBH), 2,2-bis(4-chlorophenyl) ethanol (DDOH), 1-chloro-2,2-bis (4-chlorophenyl) ethylene (DDMU), bis(4-chlorophenyl) ketone (dichlorobenzophenone) (4,4'-DCBP), 4-chlorobenzoic acid (4-CBA), 1,1-dichloro-2,2-bis(4-sulfophenyl)ethane (DDMS), Hexachlorocyclohexane (HCH), Spent Mushroom Substrate (SMS), Tetrachlorocyclohexene (TCCH), Tetrachlorocyclohexanol (TCCOL), Tetrachlorocyclohexene epoxide (TCCE), Veratryl Alcohol (VA), Manganese Peroxidase (Mg-P), Lignin Peroxidase (Li-P), Pentachlorocyclohexanol (PCCOL), Tetrachlorocyclohexanediol (TCCDL), Trihydroxytrichlorocyclohexane (THTCH), Diacetylated trichlorocyclohexanediol (Tri-CCDL), 1,3,4,6, -tetrachloro-1,4-cyclohexadiene(1,4-TCCHdiene), 1,4-dichlorobenzene (1,4 DCB), 1,2,4-trichlorobenzene (1,2,4 TCB), chloro-cis-1,2-dihydrocyclohexadiene (CDCHdiene), maleylacetate (MA), and 3-chlorocatechol (3-CC), γ-pentachlorocyclohexane (γ-PCCH), Trichloro-2,5-cyclohexadiene-1-ol (TCCH-1-ol), 2,5-Dichlorohydroquinone (2,5-DCHQ), 2-Chlorohydroquinone (2-CHQ), Hydroxymuconicsemialdehyde (HMSA), Pentachloroanisole (PCA), Tetrachloroanisole (TeCA), Tetrachlorophenol (TeCP), Tetrachlorohydroquinone (TCHQ), Chlorohydroquinone (CHQ), 2,6-Dichloro-1,4-benzoquinone (DCBQ), 2,3,4,6-tetrachloroanisole (2,3,4,6-TeCA), Tetrachloro-4-methoxyphenol (TCMP), tetrachloro-1,4-dimethoxybenzene (TCDB).
anaerobic digester sludge, rich in methanogens, provides an abundant and economical source for such applications (Jablonski et al., 1996). Furthermore, Orndorfff and Colwell (1980) studied the microbial transformation of chlordecone by *Pseudomonas aeruginosa* strain KO3. Their results showed that this strain was able to partially transform chlordecone into monohydrochlordecone (15 %) and dihydrochlordecone (5 %) after an eight-week incubation under aerobic conditions. However, this work has been criticized because the products detected could be impurities initially present in the chlordecone used, rather than biodegradation products. Furthermore, none of these bacterial strains were able to use chlordecone as a sole carbon source, suggesting that complete mineralization of chlordecone by bacteria is unlikely (Orndorfff and Colwell, 1980). Recent research has highlighted the ability of *Desulfovibrio* sp.86 to transform chlordecone (Della-Negra et al., 2020). Under anaerobic conditions, this bacterial strain performs dechlorination with opening of the chlordecone cycle, producing hydrochlordecones, polychloroindenes and polychloroindenecarboxylics. Notably, by modifying the operating conditions, Desulfovibrio sp.86 initiates a reductive sulfidation, leading to the formation of a novel thiolated derivative of chlordecone. These results, combined with studies on *Methanosarcina thermophila* and *Pseudomonas aeruginosa* strain KO3, highlight the diversity of microbial mechanisms involved in the transformation of chlordecone, from anaerobic dechlorination to reductive sulfidation. However, the persistence of this pollutant in the environment highlights the importance of exploring other approaches, including the involvement of fungi, whose potential in the bioremediation of organochlorine compounds remains underexploited.

### 3.2.4. Lindane

Hexachlorocyclohexane (HCH) is a cyclic, polychlorinated hydrocarbon existing in nine stereoisomeric forms, among which α-, β-, γ-, and δ-HCH are of primary importance. Technical lindane consists of a mixture of several hexachlorocyclohexane (HCH) isomers manufactured through the photochemical chlorination of benzene (Jackovitz and Hebert, 2015). Among HCH isomers, γ-HCH commonly known as lindane is the only isomer exhibiting effective insecticidal properties. Thereby, lindane has been extensively used for the control of agricultural pests along with its use as a scabicidal and pediculicidal agent to combat human and veterinary medical pests. Although the widespread use of technical-grade HCH and lindane has been discontinued and severely restricted for several decades, as a consequence of its toxicity, its long-range transport capacity, and its hydrophobic nature conferring its persistence and ability to bioaccumulate and biomagnify in the food chain, the detection of lindane residues in environmental matrices and at different trophic levels remains problematic to this date (Guillén-Jiménez et al., 2012).

Microbial remediation of lindane-polluted environments has been largely investigated during the past few decades as an alternative to demanding technologies that are currently employed such as thermal desorption and incineration. Previous studies addressing microbial remediation of lindane have reported the degradation ability of numerous fungal species including *Phanerochaete chrysosporium* (Mougin et al., 1996; Merlin et al., 2014; Singh and Kuhad, 1999), *Trametes hirsutus* (Singh and Kuhad, 1999), *Cyathus bulleri* (Singh and Kuhad, 2000), *Phanerochaete sordida* (Singh and Kuhad, 2000), *Pleurotus ostreatus* (Rigas et al., 2005; Cigić et al. 2012; Ulčnik et al., 2013), *Ganoderma australe* (Rigas et al., 2007), *Conidiobolus* sp. (Nagpal et al., 2008), *Bjerkandera adusta* (Quintero et al., 2008), *Pleurotus eryngii* (Quintero et al., 2008), *Polyporus ciliatus* (Quintero et al., 2008), *Irpex lacteus* (Quintero et al., 2008), *Fusarium solani* (Sagar and Singh, 2011), *Fusarium poae* (Sagar and Singh, 2011), *Hypoxylon fragiforme* (Cigić et al. 2012), *Trametes versicolor* (Cigić et al. 2012; Ulčnik et al., 2013), *Rhodotorula* sp. VITJzN03 (Abdul Salam et al., 2012), *Ganoderma lucidum* (Kaur et al., 2016), *Fusarium verticillioides* (Guillén-Jiménez et al., 2012), *Phlebia lindtneri* (Xiao and Kondo 2020a), *Phlebia brevispora* (Xiao and Kondo 2020a), *Candida* VITJzN04 (Salam and Das, 2014), *Pleurotus florida* (Mohapatra et al. 2015) Detailed data on the degradation of this pollutant are summarized in Table 1 (section color code: grey). This highlights the fungal species that have demonstrated biodegradation potential, specifying the degradation rates obtained, the experimental conditions (incubation time and temperature), the culture medium or the parameters of the system used, as well as the degradation products identified and the associated references. Moreover, degradation pathways employed by fungal species involved in the transformation of Lindane are depicted in Supplementary Figure 6.

### 3.2.5. Pentachlorophenol

Pentachlorophenol (PCP) is a widespread persistent environmental pollutant that has been, and in some countries remains, one of the most extensively used pesticides, especially in wood preservation as a fungicidal agent (Sing et al., 2014). From a structural point of view, PCP is a penta-chlorinated derivative of phenol. Owing to its extreme persistence in soil and aquatic environments, the U.S EPA has classified PCP as a priority pollutant, hence the constant need for eco-friendly, innovative technologies allowing the removal of this pesticide from various environmental matrices, and soils in particular.

Fungi have emerged as a promising option for the remediation of PCP-contaminated sites. The ability to degrade pentachlorophenol has been observed among species belonging to the phyla Basidiomycota (Boyle, 2006; Cea et al., 2010; Chung and Aust, 1995; Okeke et al., 1997; Rubilar et al., 2007; Tortella et al., 2008; Tuomela et al., 1998; Walter et al., 2005b; Walter et al., 2005a; Souza et al., 2011; Rubilar et al., 2011; Xiao and Kondo 2020b), Ascomycota (Vacondio et al., 2015; Bosso et al., 2015; 2016; Hechmi et al., 2016; Carvalho et al., 2009), and Mucoromycota (Carvalho et al., 2011; Carvalho et al., 2013; Sing et al., 2014; Ruiz-Lara et al., 2020), (Table 1, Supplementary Figure 7).

Worth noting, PCP has been reported to inhibit fungal growth germination, therefore the usage of spores as inoculum in degradation or tolerance assays has been found to be inexpedient. However, when fungal mycelia are used as inoculum, sufficient fungal biomass was obtained shortly afterward (Carvalho et al., 2011). Moreover, the application of pentachlorophenol to soil matrices has been shown to exert a drastic effect on microbial activity, elucidated by a decrease in soil enzymatic indicators such as dehydrogenases and fluorescein diacetate hydrolases (Bosso et al., 2015). For more details on the degradation of Pentachlorophenol, we have grouped the information in Table 1 (section color code: pale peach), which lists the fungal species involved, their efficiency in terms of
percentage of degradation, the specific incubation conditions, as well as the final products identified and the corresponding references. In addition, degradation pathways utilized by fungal species involved in the transformation of PCP are resumed in Supplementary Figure 7.

# 4. Fungal bioremediation of industrial or unintentional persistent organic pollutants: close-up on species and capabilities

## 4.1. Polychlorinated dibenzo-p-dioxins (PCDDs) and polychlorinated dibenzo-p-furans (PCDFs)

Polychlorinated dibenzo-p-dioxins (PCDDs) and Polychlorinated dibenzo-p-furans (PCDFs) are tricyclic aromatic compounds harboring similar physicochemical properties and comprising, from a structural point of view, 1-8 chlorine atoms occurring at various positions. 135 and 175 congeners of PCDDs and PCDFs exist, respectively, as a result of the number and permutations of the chlorine atoms. In addition, the number of chlorine atoms along with their position on the skeleton affects the toxicity and solubility of these compounds with 2,3,7,8-tetrachlorodibenzo-p-dioxin (2,3,7,8-tetraCDD), being the most toxic among the PCDD/PCDFs congeners ("Chapter 5.11 Polychlorinated Dibenzodioxins and Dibenzofurans (n.d.)"). Dibenzo-p-dioxins with an increased number of lateral chlorines have been considered more toxic, compared to congeners with decreased number of lateral chlorines or those with an increased number of non-lateral chlorine atoms ("Chapter 5.11 Polychlorinated Dibenzodioxins and Dibenzofurans (n.d.)"). These chloroaromatic compounds are generated as undesirable impurities resulting from the incomplete combustion of chlorinated chemicals, municipal solid wastes, and from industrial manufacture of herbicides and bleach paper pulp employing chlorinated compounds (Hiraishi, 2003).

Actual removal techniques include namely thermal desorption or high-temperature incineration (> 1100 °C). Although efficient, these physicochemical techniques are high-cost and can only be applied to relatively small amounts of contaminated soil at a time. Therefore, from a sustainable point of view, the application of microbial consortia capable of remediating dioxin- and furan-polluted environments has a greater appeal compared to physicochemical methods currently employed (Hiraishi, 2003).

While the number of isolated fungal species capable of degrading 2,3,7,8-tetraCDD is still rather low, more fungi have been reported to display the capacity to break down other polychlorinated dibenzo-dioxin congeners. Detailed information on the biodegradation of PCDDs/PCDFs is provided in Table 2 (section color code: light green). This table lists the fungal species that have demonstrated the ability to degrade it, specifying the degradation percentages obtained, the incubation conditions (duration and temperature), the type of culture medium used, as well as the degradation products detected and the references used.

## 4.2. Polychlorinated biphenyls (PCBs)

Polychlorinated biphenyls are a class of synthetic organic compounds containing a biphenyl core in which hydrogen atoms are substituted by chlorine atoms at any of the 10 possible positions. Thus, depending on the number of chlorine atoms and their location on the biphenyl ring, 209 different compounds referred to as congeners can arise from such substitutions (Mouhamadou et al., 2013). PCBs have been largely used in electrical equipment, as hydraulic fluids, plasticizers, or as fire retardants due to their desirable physical properties including their thermal and chemical stability, and their low electrical conductivity (Mouhamadou et al., 2013). Similarly, to other persistent organic pollutants, PCBs can be readily accumulated in fatty tissues through their lipophilicity, thus exerting their toxicity and potential carcinogenicity. As a consequence of their toxicity and their deleterious environmental effects, PCBs are ranked 5th on the ATSDR Priority List of Hazardous Substances (Agency for Toxic Substances and Disease Registry (ATSDR), n.d.).

Although several physicochemical technologies are at hand for the cleanup of PCB-contaminated matrices, bioremediation technologies employing microorganisms as remediation agents offer a complementary or even an alternative strategy with lower environmental impact allowing preservation of environmental matrices along with their inhabitants.

Bacterial remediation of PCBs has proven unsatisfactory due to inherent limitations of the microbial mechanisms involved. Indeed, PCB degradation relies mainly on the action of biphenyl catabolic enzymes, but this process results in the accumulation of chlorobenzoic acids (CBAs), considered as dead metabolites that are not easily degradable and persist in the environment, thus reducing the overall efficiency of the process (Field and Sierra-Alvarez, 2008). Moreover, another major constraint of bacterial bioremediation of PCBs lies in the intracellular nature of the degradation process. To be degraded, the pollutant must first penetrate inside bacterial cells, which is a rate-limiting step, particularly for hydrophobic compounds such as PCBs. This barrier to intracellular uptake significantly reduces the efficiency of bacterial biodegradation (Cvančarová et al. 2012). As opposed to bacterial degradation systems, in which the extent of PCBs degradation is affected by the number and position of chlorine substitutes, fungal degradation of PCBs appears to occur with no selectivity, as congeners with different degrees of chlorination degrade to similar extents by both model white-rot fungi (Federici et al., 2012; Gayosso-Canales et al., 2012), as well as autochthonous species (Mouhamadou et al., 2013; Sage et al., 2014; Germain et al., 2021; J. Germain et al., 2021a). Table 2 (section color code: light grey) provides a more in-depth overview of the biodegradation processes associated with PCBs. It details the fungal species involved, their degradation performances, the experimental conditions implemented (time, temperature and culture medium), as well as the main transformation products identified and the corresponding bibliographic sources.

## 4.3. Polybrominated diphenyl ethers (PBDEs)

Polybrominated diphenyl ethers (PBDEs) are a group of emerging organobromine pollutants belonging to the brominated flame
<caption>**Table 2**</caption>

Primary fungal species capable of degrading the three industrial and By-products POPs (Polychlorinated dibenzo-p-dioxins (PCDDs) and Polychlorinated dibenzo-p-furans (PCDFs), Polychlorinated Biphenyls PCBs, and Polybrominated Diphenyl Ethers (PBDEs)). Color coding system: PCDDs and PCDFs (light green), PCBs (light grey), PBDEs (orange).

<table><thead><tr><th>Fungal Species</th><th>Degradation %</th><th>Incubation</th><th>Medium</th><th>Degradation Products</th><th>Notes</th><th>Reference</th></tr></thead><tbody><tr><td>Phanerochaete chrysosporium<br>OGC101</td><td>50%</td><td>27d, 38°C,<br>pH=4.5</td><td>Nitrogen-limited<br>(1.2 mM ammonium<br>tartrate) liquid cultures</td><td>1,2,4-THB, 1-chloro-3,4-dihydroxybenzene, 2-<br>methoxy-1,4-dihydroxybenzene, 4-chloro-1,2-<br>benzoquinone, 1-chloro-3,4-methoxybenzene,<br>2-hydroxy-1,4-benzoquinone, 2-methoxy-1,4-<br>benzoquinone, and chloride</td><td>Involvement of Li-P, Mn-P, and<br>methyltransferases<br>2,7-Dichlorodibenzo-p-Dioxin (25 µM)<br>A multistep degradation pathway was suggested</td><td>(Valli et al.,<br>1992)</td></tr><tr><td>Phanerochaete sordida<br>YK-624</td><td>40% (Tetrachloro-) - 76%<br>(Hexachloro-) PCDDs,<br>45% (Tetrachloro-) -<br>70% (Hexachloro-) PCDFs,</td><td>14d, 30°C,<br>pH=4.5</td><td>Stationary low-nitrogen<br>(LN) basal III liquid<br>medium</td><td>4,5-dichloratecolch,<br>and Tetrachloratecolch<br>NM</td><td>Capable of degrading a mixture of ten congeners of<br>PCDDs and PCDFs (tetra- to octa-PCDDs and<br>PCDFs), conc 0.05-mg/ml (500 µg each)</td><td>(Takada et al.,<br>1996)</td></tr><tr><td>Phlebia lindneri</td><td>250 nmol → 149 nmol of CO<sub>2</sub></td><td></td><td></td><td></td><td>Tetrachlorinated congeners of dibenzo-p-dioxins<br>and dibenzo-p-furans endured the lowest<br>degradation</td><td>(Takada et al.,<br>1996)</td></tr><tr><td>Phlebia sp. MG-60</td><td>250 nmol → 155 nmol of CO<sub>2</sub></td><td>30d, 30°C,<br>pH=4.5</td><td>Low-nitrogen (LN)<br>basal III liquid medium</td><td></td><td>Hexachlorinated congeners of these compounds<br>underwent the highest degradation rates</td><td></td></tr><tr><td>Unidentified strain MZ-227</td><td>250 nmol → 196 nmol of CO<sub>2</sub></td><td></td><td></td><td></td><td>High concentrations (100 µM) of 2,7-<br>Dichlorodibenzo-p-Dioxin</td><td>(Mori and<br>Kondo, 2002a)</td></tr><tr><td>Phlebia lindneri<br>strain GB-1027</td><td>16.5% (2,7-diCDD),<br>35% (after 20d)<br>5.8% (2,8-diCDF)</td><td>5-20d, 30°C,<br>pH=4.5</td><td>Low-nitrogen (LN)<br>basal III liquid medium</td><td>Hydroxy-diCDD<br>Hydroxy-diCDF<br>2,7-DiCDD:<br>Monohydroxy-diCDD<br>Monomethoxy-diCDD<br>Dimethoxy-diCDD</td><td>Hexachlorinated congeners of these compounds<br>underwent the highest degradation rates</td><td>(Mori and<br>Kondo, 2002b)</td></tr><tr><td>Phlebia lindneri</td><td>30% (2,7-diCDD and<br>1,2,6,7-tetraCDD),<br>10% (2,3,7-triCDD and<br>1,2,8,9-tetraCDD),<br>0% (1,3,6,8-tetraCDD)</td><td>5-21d, 30°C,<br>pH=4.5</td><td>Low-nitrogen (LN)<br>basal III liquid medium</td><td>2,3,7-TriCDD:<br>Monohydroxy-triCDD<br>Monomethoxy-triCDD<br>1,2,6,7-TetraCDD:<br>Monohydroxy-tetraCDD<br>Monomethoxy-tetraCDD<br>1,2,8,9-TetraCDD:<br>Monohydroxy-tetraCDD<br>Monomethoxy-tetraCDD</td><td>Transformation of 2,7-diCDD, 2,3,7-triCDD,<br>1,2,6,7-, 1,2,8,9-, and 1,3,6,8-tetraCDD (25 µM)<br>was investigated<br>Hydroxylation followed by further methoxylation<br>> major degradation pathway<br>2,7-diCDD and 1,2,6,7-tetraCDD endured the<br>highest degradation<br>No degradation of 1,3,6,8-tetraCDD was observed</td><td>(Kamei et al.,<br>2005)</td></tr><tr><td>Panellus sticticus<br>strain 99-334</td><td>≈ 100%</td><td>15-40d,<br>25°C</td><td>Low-nitrogen (LN)<br>liquid medium</td><td>1-chloro-3,4-dihydroxybenzene</td><td>2,7-DCDD (10 µM) was investigated,<br>Presumable involvement of the intracellular<br>cytochrome P450 monooxygenase rather than<br>ligninolytic enzymes, albeit being a WRF</td><td>(Sato et al.,<br>2002)</td></tr><tr><td>Phlebia brevispora<br>TMIC33929</td><td>80% (2,7-diCDD)<br>72% (1,2,6,7-tetraCDD)<br>55% (2,3,7-triCDD)<br>37% (1,2,8,9-tetraCDD)<br>≈ 50% (1,3,6,8-tetraCDD, in<br>the historically polluted soil)<br>≈ 80% (1,3,6,8-tetraCDD, soil<br>microcosm under slurry state,<br>solid:liquid ration is 1:60,<br>under 30d)<br>≈ 50% (2,7-DCDD, soil<br>microcosm, organic-poor soil)</td><td>28d, 30°C,<br>pH=4.5</td><td>Low-nitrogen (LN)<br>liquid medium</td><td>Monohydroxy-tetraCDD,<br>Monomethoxy-tetraCDD,<br>DimethoxytetraCDD,<br>3,5- Dichloratecolch,<br>and Dimethoxy-triCDD</td><td>Degradation of the recalcitrant dioxin congener<br>1,3,6,8-tetraCDD (0.25 µmol/flask) was first<br>observed with this strain<br>Conc of all PCDDs (0.25 µmol/flask)<br>Transformation appeared to occur through<br>hydroxylation and subsequent methoxylation<br>reactions,<br>Presumable involvement of a cytochrome P450<br>monooxygenase</td><td>(Kamei et al.,<br>2005)</td></tr><tr><td>Phlebia brevispora<br>TMIC33929</td><td>≈ 50% (1,3,6,8-tetraCDD, in<br>the historically polluted soil)<br>≈ 80% (1,3,6,8-tetraCDD, soil<br>microcosm under slurry state,<br>solid:liquid ration is 1:60,<br>under 30d)<br>≈ 50% (2,7-DCDD, soil<br>microcosm, organic-poor soil)</td><td>5-90d,<br>25°C</td><td>Three types of soil<br>microcosms (sterilized<br>soils)<br>Historically polluted<br>paddy soil under<br>slurry-state conditions</td><td>NM</td><td>1,3,6,8-tetraCDD (0.05 mM)<br>The presence of organic materials in soils reduces<br>1,3,6,8-TCDD degradation<br>Water content → Crucial factor in the remediation<br>of contaminated soils</td><td>(Kamei et al.,<br>2009)</td></tr><tr><td>Coprinellus dixidentatus<br>TUFC11148 and<br>TUFC34534</td><td>98.7% (DD)<br>18.2% (2,7-DCDD)</td><td></td><td></td><td>Monohydroxylated 2,7-DCDD</td><td>DD (1 µmol), 2,7-DCDD (0.1 µmol)<br>Presumable involvement of the intracellular<br>cytochrome P450 monooxygenase system<br>Contrarily to other Basidiomycota, Coprinellus<br>species are known to inhabit and tolerate soil<br>environments → Advantage as soil bioremediation<br>agents</td><td>(Suhara et al.,<br>2011)</td></tr><tr><td>Coprinellus micaceus<br>TUFC30081</td><td>61.6% (DD)<br>51.8% (2,7-DCDD)</td><td>7-14d, 25°C,<br>pH=5.6</td><td>High-carbon-low-<br>nitrogen (HCLN)<br>liquid Kirk medium</td><td>Monohydroxylated 2,7-DCDD</td><td>WHO-TEQ value was 72960 ± 13400 ng kg<sup>-1</sup><br>before treatment, with 1,2,3,4,6,7,8-HpCDF as the<br>most abundant congener with 663330 ± 1 167 620<br>ng kg<sup>-1</sup></td><td>(Anasonye et<br>al., 2014)</td></tr><tr><td>Phanerochaete velutina<br>FBCC nb 941</td><td>62% (WHO-TEQ-value)</td><td></td><td>3 months,<br>at room T°C,<br>pH=6.1</td><td>NM</td><td>Pine bark was used as substrate to introduce fungi<br>to soils<br>No degradation observed using fungal extracellular<br>enzymes</td><td>(Anasonye et<br>al., 2014)</td></tr><tr><td>Stropharia rugosoanulata<br>FBCC nb 475</td><td>64% (WHO-TEQ-value)</td><td></td><td>Non-sterile soil<br>microcosms from<br>historically<br>contaminated sawmill<br>areas</td><td>NM</td><td>Contrarily to other Basidiomycota, Coprinellus<br>species are known to inhabit and tolerate soil<br>environments → Advantage as soil bioremediation<br>agents</td><td>(Anasonye et<br>al., 2014)</td></tr><tr><td>Phanerochaete velutina<br>FBCC nb 941</td><td>62% (WHO-TEQ-value)</td><td></td><td>3 months,<br>at room T°C,<br>pH=6.1</td><td>NM</td><td>Contrarily to other Basidiomycota, Coprinellus<br>species are known to inhabit and tolerate soil<br>environments → Advantage as soil bioremediation<br>agents</td><td>(Anasonye et<br>al., 2014)</td></tr><tr><td>Cordycops militaris KS-92</td><td>55% (2,7-diCDD),<br>47% (2,3,7-triCDD),<br>24% (2,4,8-triCDF)</td><td>30d, 25°C</td><td>Liquid Glucose-BSM<br>Medium</td><td>2,7-diCDD:Hydroxyl-diCDD, Dihydroxyl-<br>diCDD<br>2,3,7-triCDD:Hydroxyl-triCDD<br>2,4,8-triCDF:Hydroxyl-triCDFs</td><td>Conc of all PCDDs (0.25 µmol/flask)<br>Central role for the CYP450 monooxygenase in the<br>degradation of chlorinated dioxins</td><td>(Mori et al.,<br>2015)</td></tr></tbody></table>

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<table><caption>Table 2 (continued)</caption><thead><tr><th></th><th></th><th></th><th></th><th></th><th></th><th></th></tr></thead><tbody><tr><td>Cordyceps sinensis<br>strain A</td><td>~ 58% (DD)<br>~ 48.6% (2,3,7-triCDD)<br>~ 59% (octaCDD)</td><td>7d,<br>30°C,<br>pH=6.8</td><td>Basal Salt liquid<br>Medium (BSM)</td><td>Catechols<br>Trichloro cis,cis-macronates (from octaCDD),<br>Dichloro cis,cis-macronates (from 2,3,7-<br>triCDD)</td><td>Degradation of DD, 2,3,7-triCDD, and octaCDD<br>was investigated (0.003%)</td><td></td></tr><tr><td>Rigidoporus sp. FMD21</td><td>73%</td><td>28d, 30°C,<br>pH=7</td><td>PDSRB (Potato Dextrose<br>Soy Rice Ijran) liquid<br>medium</td><td>NM</td><td>Hydrolysis reactions were not employed in the<br>degradation of dibenz-p-dioxins<br>Chlorine atoms were replaced by hydrogens rather<br>than hydroxyl groups<br>Degradation of 2,3,7,8-tetraCDD (0.5 pg TEQ/µL<br>TCDD) was investigated<br>This strain displayed the ability to secrete<br>substantial amounts of laccase and Mn-P</td><td>(Nakamiya et al., 2005)<br>(Dao et al., 2019)</td></tr><tr><td>Pseudallescheria boydii</td><td>T<sub>1</sub> = 0.57 ng<br>TCDD/ing cell<br>when C<sub>1</sub>=4.4<br>ng/mL</td><td>30°C</td><td>Normal Medium<br>(Liquid)</td><td>NM</td><td>Degradation of 2,3,7,8-tetraCDD (50 ng) was<br>investigated<br>Transformation occurred through co-metabolism<br>using glucose as a carbon source<br>Degradation occurred in the log phase of P. boydii</td><td>(Ishii et al., 2009)</td></tr><tr><td>Rigidoporus sp. FMD21<br>(Laccases)</td><td>50%</td><td>14d,<br>37°C,<br>pH=3.7</td><td>Buffered Minimal<br>Methanol medium<br>(BMM),<br>Buffered methanol-<br>complex medium<br>(BMMY),<br>Laccase producing<br>medium</td><td>NM</td><td>Transformation of 2,3,7,8-TCDD (0.05 pgTEQ/µL)<br>by recombinant laccase was investigated<br>The genome of Rigidoporus sp. FMD21 was<br>sequenced and investigated for LME-encoding<br>genes<br>Seven of the predicted laccase genes were then<br>subjected to heterologous overexpression in Pichia<br>Pastoris X-33</td><td>(Dao et al., 2021b)</td></tr><tr><td>Rigidoporus sp. FMD21<br>(Crude Extracellular Enzyme<br>Extract or ExE)</td><td>77%</td><td>36d, 30°C</td><td>PDSRB Liquid Medium</td><td>3,4-dichlorophenol</td><td>Four of the seven recombinant laccases revealed<br>important activity towards ABTS, in addition to<br>the capacity to bind and potentially oxidize 2,3,7,8-<br>tetraCDD during in-silico dockings, + concordance<br>with the in-vitro results<br>Transformation of 2,3,7,8-TCDD (0.5 pgTEQ/µL)<br>was investigated<br>The starting laccase activity in the ExE was 80,000<br>U/L</td><td>(Dao et al., 2021a)</td></tr><tr><td>Pleurotus pulmonarius<br>strain BCRC36906</td><td>~ 100% (TetraCDD/Fs,<br>PentaCDD/Fs)<br>>80% (HexaCDD/Fs)<br>>97% (HeptaCDD/Fs)<br>>90% (OctaCDD/Fs)<br>96% (Overall PCDD/F<br>Removal)</td><td>72d,<br>26 ± 2 °C</td><td>Solid State Fermentation<br>(SSF) of field soil,<br>Non-sterilized soil with<br>a total PCDD/F conc of<br>14,000 ± 2,400 ng I-<br>TEQ kg<sup>-1</sup></td><td>NM</td><td>Degradation did not significantly vary with<br>2,3,7,8-TCDD concentrations ranging between<br>0.005 and 0.5 pgTEQ/µL<br>The addition of violuric acid (0.5 mM), a synthetic<br>laccase mediator, significantly increased 2,3,7,8-<br>tetraCDD degradation to an additional 23% rate<br>Laccase and Mn-P detected in this fungal strain<br>Only Mn-P displayed a strong positive correlation<br>with PCDD/Fs degradation<br>PCDD/Fs removal revealed no congener-specificity<br>Increasing the inocula size (1:1.4 w/w ratio)<br>improved the degradation rate of PCDD/Fs through<br>with P. pulmonarius.<br>Residual conc of PCDD/Fs after removal was<br>Below the regulatory control limit (1,000 ng I-<br>TEQ kg<sup>-1</sup>)</td><td>(Kawelavvoong et al., 2020)</td></tr><tr><td>Penicillium sp. strain Q1-1</td><td>87.90%</td><td>6d, 31°C,<br>pH=7.4</td><td>Potato Dextrose Broth<br>(PDB)</td><td>4,5-Dichloro-1,2-benzoquinone,<br>2-Hydroxy-1,4-benzoquinone,<br>4,5-Dichlorocatechol,<br>1,2,4-Trihydroxybenzene,<br>β-ketoacidic acid</td><td>Degradation of 2,3,7,8-TCDD (1 µg/mL) was<br>investigated<br>The biodegradation kinetics were fitted to a first-<br>order kinetic model<br>None of the degradation products/metabolites<br>persisted at the end of the experiment</td><td>(Qiu et al., 2021)</td></tr><tr><td>Fusarium solani</td><td>24.7% (PCDD/Fs)<br>Reduced Soil's TEQ by 26.6%</td><td>2 months,<br>20 ± 1 °C</td><td>Aged Dioxin-<br>Contaminated<br>Soil Microcosms<br>(Non-Sterile), (260.09 ±<br>37.10 ng kg<sup>-1</sup>)</td><td>NM</td><td>The addition of cardboard<br>chips rather than wood<br>chips as a bulking agent,<br>and olive oil as a carbon<br>source -> better<br>PCDD/Fs dissipation and<br>reduction of the toxicity equivalence (TEQ)</td><td>(Delsarte et al., 2021)</td></tr><tr><td>Penicillium brasilianum</td><td>15.8% (PCDD/Fs)<br>Reduced Soil's TEQ by 35%</td><td></td><td></td><td></td><td>Degradation of a technical mixture of PCBs (200<br>µg of Delor 103 / flask) was investigated<br>All eight ligninolytic fungal species tested in this<br>study significantly degraded the pollutant with P.<br>ostreatus being by far the most effective degrader<br>The ability to suppress the associated toxicity<br>through ecotoxicological assays employing the<br>luminescent bacterium Vibrio fischeri<br>Only P. ostreatus suppressed PCBs-associated<br>toxicity for the entire incubation period<br>Both the intracellular CYP450 monooxygenase<br>system and LMEs seem to be involved in the<br>degradation of PCBs</td><td>(Čvančárović et al., 2012)</td></tr><tr><td>Pleurotus ostreatus<br>3004 CCBAS 278</td><td>99.6% and 98.4% (Malt-Extract<br>Glucose Medium)</td><td>28-42d,<br>28°C</td><td>Low Nitrogen Mineral<br>Medium (LNMM), static<br>cultures</td><td>Forty-one PCB degradation products<br>including: D<sub>5</sub>, tri- and tetrachlorinated<br>derivatives of hydroxy and methoxy biphenyls,<br>chlorobenzonic acids, chlorinated<br>Benzaldehydes, and chlorinated alcohols</td><td></td><td></td></tr></tbody></table>

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<table><caption>Table 2 (continued)</caption><thead><tr><th></th><th></th><th></th><th></th><th></th><th></th></tr></thead><tbody><tr><td>Pleurotus sajor-caju LBM 105</td><td>91.7% and 97.7% (Complex Medium)</td><td>35d, 29 ± 1 °C</td><td>N-limited Liquid Mineral Medium (GA)</td><td>NM</td><td>Degradation of technical mixture of PCBs (Aroclor 1242, 1254 y 1260, at 217 mg L<sup>-1</sup>) was investigated<br>PCB removal was accompanied by a reduction in PCB-related toxicity as the radial growth of Ganoderma sp., a toxicity indicator, increased following treatment with P. sajor-caju.<br>The addition of PCBs increased proteins and laccase secretion, changes in biomass, hyphal diameter, and glucose consumption<br>Mycoremediation of PCB using GA costed 0.69 $ g<sup>-1</sup> d<sup>-1</sup></td><td>(Sadañoski et al., 2019)</td></tr><tr><td>Co-Cultures of Pleurotus sajor-caju LBM 105 and Trametes sanguinea LBM 023</td><td>91%</td><td>90d, 25 ± 1 °C</td><td>Microcosms experiments using Ultisol soil (non-sterile)</td><td>NM</td><td>Degradation of technical mixture of PCBs (Aroclor 1242, 1254 and 1260, at 69.547 ± 9.799 µg g<sup>-1</sup>) was investigated<br>Mycoaugmentation with co-cultures of P. sajor-caju and T. sanguinea immobilized on sugarcane bagasse was tested<br>A significant toxicity attenuation was observed after 90d<br>Tested strains were capable of surviving in non-sterilized soils<br>Biostimulation with sugarcane bagasse resulted in similar removal rates (88%), whereas natural Attenuation was deemed inefficient</td><td>(Sadañoski et al., 2020)</td></tr><tr><td>Pleurotus pulmonarius LBM 105</td><td>95.40%</td><td></td><td></td><td></td><td>Degradation of technical mixture of PCBs (Aroclor 1242, 1254 y 1260, at 217 mg L<sup>-1</sup>) was investigated<br>Higher PCB removal rates were observed when with monocultures of P. pulmonarius and T. sanguinea compared with co-cultures of the two strains (43%)<br>LMEs secretion profile changed in co-cultures</td><td>(Benitez et al., 2021)</td></tr><tr><td>Trametes sanguinea LBM 023</td><td>55%</td><td>24d, 28 ± 1 °C</td><td>N-limited Liquid Mineral Medium (GA)</td><td>NM</td><td>Toxicity of intermediate products was lower than that of parental pollutant<br>Proteomic analysis revealed a transcriptional response to the PCB-induced oxidative stress through up-regulation of the glyceraldehyde-3-phosphate dehydrogenase (632) and the basal transcription factor 3, in addition to an increase in the production of short-chain and aldol/keto reductases, laccases and peroxidases</td><td>(Sebastian et al., 2021)</td></tr><tr><td>Pleurotus pulmonarius LBM 105</td><td>65.50 ± 8.09%</td><td>28d, 28 ± 1 °C</td><td>N-limited Liquid Mineral Medium (GA)</td><td>NM</td><td>Exposure to PCBs on the other hand induced the down-regulation of several structural chaperone and transportation proteins including actin, β-tubulin, and heat shock proteins<br>A 46.47% toxicity reduction was observed</td><td>(Sebastian et al., 2021)</td></tr><tr><td>Leninus tigrinus CBS 577.79</td><td>33.6 ± 0.3%</td><td>60d, 28°C</td><td>Microcosms soil historically contaminated by Aroclor 1260 (776.3 ± 1.2 mg kg<sup>-1</sup> of soil)</td><td>NM</td><td>Addition of soybean oil as a PCB-mobilizing agent is not advisable<br>The impact of Mycoaugmentation with L. tigrinus on the native microbiota and PCB remediation was investigated<br>L. tigrinus was able to efficiently colonize and compete with the native microbiota since an increase in fungal biomass denoted by an increase in ergosterol contents was observed upon bioaugmentation<br>Except for biphenyl-degrading bacteria, bioaugmentation with L. tigrinus enhanced both bacterial and fungal biodiversity</td><td>(Federici et al., 2012)</td></tr><tr><td>Pleurotus ostreatus 3004 CCBAS 278</td><td>18.5% (Bulk), 41.3% (Topsoil), 50.5% (Rhizosphere)</td><td></td><td>Microcosms soils (non-sterilized) of different typologies:<br>Bulk soil (C<sub>r</sub>=705.95 µg/g)<br>Topsoil (C<sub>r</sub>=375.81 µg/g)<br>Rhizosphere (C<sub>r</sub>=169.36 µg/g)</td><td>Methoxylated and hydroxylated PCBs, Chlorobenzoates (CBAs), and Chlorobenzyl alcohols all at trace levels, except for 2,3,6-trichlorobenzoic acid</td><td>P. ostreatus effectively colonized the soil<br>Mycoaugmentation with P. ostreatus and L. lacteus resulted in a decrease in microbial diversity denoted by a decrease in both the Chaol and Shannon diversity indexes<br>Bioaugmentation with P. ostreatus stimulated native bacterial growth, however, a shift in community composition towards Firmicutes-bacteria at the expense of Actinobacteria and Chloroflexi was observed<br>L. lacteus on the other hand had no stimulation effect on bacterial growth</td><td>(Stella et al., 2017)</td></tr><tr><td>Irpecoccus lacteus 617/93</td><td>19.3% (Bulk), 39.4% (Topsoil), 30.3% (Rhizosphere)</td><td>12 weeks, 25°C</td><td></td><td></td><td>However, PCBs transformation by these strains led to the formation of more acutely toxic products</td><td></td></tr></tbody></table>

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<table><caption>Table 2 (continued)</caption><thead><tr><th></th><th></th><th></th><th></th><th></th><th></th><th></th></tr></thead><tbody><tr><td>Extracellular Exudate of Pleurotus ostreatus 3004<br>CCBAS 278 and Irpex lacteus<br>617/93</td><td>> 80% of most OH-PCBs<br>within one hour (P. ostreatus)</td><td>7d, 28°C,<br>pH = 4.5</td><td>Malt-extract-glucose<br>liquid medium (MEG),<br>C<sub>0</sub> = 2 µg mL<sup>-1</sup> for each<br>congener/metabolite,<br>For P. ostreatus, 450 ± 9<br>U L<sup>-1</sup> of laccase, and 30<br>± 2 U L<sup>-1</sup> of Mn-P</td><td>NM</td><td>The extracellular exudate (EE) of P. ostreatus<br>containing predominantly laccase, biotransformed<br>a large array of selected hydrolyzed PCBs,<br>especially with the addition of syringaldehyde as a<br>mediator<br>Transformation rates and extent largely varied<br>among congeners in the case of CB-OHs and CB-<br>CHOs<br>A negative correlation between removal rates<br>and the chlorination level and position (ortho-) was<br>observed<br>EE of L. lacteus (+++ Mn-P activity --- laccase) +<br>glutathione, allowed the transformation of all CB-<br>CHOs as chlorine substitutions (ortho-) had no<br>suppressive effect</td><td>(Srdlová et al.,<br>2021)</td></tr><tr><td>Pleurotus ostreatus<br>ATCC 38540</td><td>63.5 ± 2.0% (20d)<br>63.8 ± 4.6% (30d)</td><td>20-30d,<br>28°C</td><td>Liquid SM Medium<br>with PCBs<br>C<sub>0</sub> = 7.100 mg/L<br>from transformer oil</td><td>NM</td><td>A correlation between laccase production by<br>P. ostreatus and PCB removal in liquid cultures<br>was reported through transcript expression analysis<br>Laccase production by P. ostreatus increased up to<br>6.9-fold (10d) after PCBs induction</td><td>(Gayoso-<br>Canáles et al.,<br>2012)</td></tr><tr><td>Doratomycetes nanus</td><td>73.81 ± 2.78</td><td></td><td></td><td></td><td>The potential of 24 autochthonous filamentous<br>strains isolated from a PCB-contaminated<br>industrial site was investigated</td><td>(Mouhamadou<br>et al., 2013)</td></tr><tr><td>Doratomycetes purpureofasciatus</td><td>82.01 ± 3.97</td><td></td><td></td><td></td><td>The six fungal strains of interest were also non-<br>pathogenic strains, which is of great importance in<br>the context of bioremediation</td><td></td></tr><tr><td>Doratomycetes verrucisporus</td><td>70.02 ± 7.08</td><td>7d,<br>25-37°C</td><td>GS Liquid Medium +<br>glucose<br>(C<sub>0</sub> = 3 µg/mL of each<br>congener)</td><td>NM</td><td></td><td></td></tr><tr><td>Myceliophthora thermophila</td><td>74.9 ± 1.10</td><td></td><td></td><td></td><td></td><td></td></tr><tr><td>Phoma euprena</td><td>81.59 ± 1.81</td><td></td><td></td><td></td><td></td><td></td></tr><tr><td>Thermoauscus crustaceus</td><td>78.41 ± 0.74</td><td></td><td></td><td></td><td></td><td></td></tr><tr><td>Consortium of 6 strains:<br>Thermoauscus crustaceus<br>Doratomycetes nanus,<br>Doratomycetes purpureofasciatus,<br>Myceliophthora thermophila,<br>Doratomyces verrucisporus,<br>and Phoma euprena</td><td>16.9%<br>(6 indicator PCBs i-PCBs)<br>18.7% (Total 15 PCBs)</td><td>6 months,<br>at room T°C</td><td>Mesocosms experiments<br>(C<sub>0</sub> = 400 mg kg<sup>-1</sup> total<br>of 15 PCBs)</td><td>NM</td><td>Transformation affected most of the 15 PCBs<br>analyzed with no preference for congeners with<br>lower chlorination levels</td><td>(Sage et al.,<br>2014)</td></tr><tr><td>Penicillium chrysogenum</td><td>96.56%</td><td></td><td></td><td></td><td>Isolated 12 indigenous strains among which four<br>displayed PCBs removal rates > 10%</td><td>(J. Germain et<br>al., 2021b)</td></tr><tr><td>Penicillium citreosulfuratum</td><td>95.06%</td><td>5d,<br>26°C</td><td>GS Liquid Medium +<br>glucose<br>(C<sub>0</sub> = 1 µg/mL of each<br>of the 7 congeners)</td><td>NM</td><td>Only P. canescences significantly reduced the<br>PCB-related toxicity</td><td></td></tr><tr><td>Penicillium canescences</td><td>76.72%</td><td></td><td></td><td></td><td></td><td></td></tr><tr><td>Aspergillus jensenii</td><td>84.06%</td><td></td><td></td><td></td><td></td><td></td></tr><tr><td>Consortium of 4 strains:<br>Penicillium chrysogenum,<br>Penicillium citreosulfuratum,<br>Penicillium canescences,<br>Aspergillus jensenii, and<br>Acremonium sclerotigenum</td><td>29%<br>(7 indicator PCBs<br>i-PCBs)</td><td>3 months,<br>at room T°C</td><td>Mesocosms experiments<br>(C<sub>0</sub> = 850 ppm of PCBs)</td><td>NM</td><td>Fungal biomass did not increase at the end of the<br>experiment (ergosterol assay) > Can be traced to a<br>possible decrease in the growth of native strains in<br>favor one bioaugmented ones</td><td>(Joaquim<br>Germain et al.,<br>2021)</td></tr><tr><td>Fusarium solani</td><td>84.79% - 88.11%</td><td></td><td></td><td></td><td>For each of the five fungal species, strains from<br>different origins (unpoluted UP and PCB-polluted<br>substrates) were compared for their ability to<br>degrade PCBs</td><td></td></tr><tr><td>Thermothelomyces<br>heterothallica</td><td>79.86% - 86.61%</td><td></td><td></td><td></td><td>The origin of the fungal strain has been shown<br>to have no significant effect on the tolerance<br>and resistance of the fungal species to PCBs<br>as well as their ability to degrade PCBs,<br>thus ruling out the need for a prior adaptation step</td><td>(Périgon et al.,<br>2019)</td></tr><tr><td>Thermoauscus crustaceus</td><td>95.22% - 95.91%</td><td>5d,<br>25-37°C</td><td>GS Liquid Medium +<br>glucose<br>(C<sub>0</sub> = 3 µg/mL of each<br>congener)</td><td>NM</td><td>No correlation between laccase production and<br>PCB removal was observed in this study</td><td></td></tr><tr><td>Thermothelomyces thermophila</td><td>93.44% - 94.64%</td><td></td><td></td><td></td><td></td><td></td></tr><tr><td>Schizophyllum commune</td><td>80.91% - 86.20%</td><td></td><td></td><td></td><td></td><td></td></tr><tr><td>Trametes versicolor<br>ATCC 42530</td><td>16 to 53%</td><td>26d,<br>25°C<br>pH = 4.5 ± 0.3</td><td>Sludge-bioslurry reactor<br>(sterilized sewage ludge)</td><td>NM</td><td>The degradation rates and extent with T. versicolor<br>showed no correlation with the degree of<br>bromination<br>Final sludge exhibited elevated toxicity following<br>treatment<br>Total recalcitrance was only observed with hepta-<br>BDE-183, while hexa-BDE-154 was the most<br>degradable</td><td>(Hundt et al.,<br>1999)<br>(Rodríguez-<br>Rodríguez et<br>al., 2012)</td></tr><tr><td>Unknown WRF</td><td>52.65% (native soil)<br>44.8% (sterilized soil)</td><td>4 weeks,<br>at room<br>T°C</td><td>Soil Microcosm<br>(native and<br>sterilized conditions)</td><td>NM</td><td>Biodegradation of deca-BDE (BDE-209, at<br>0.4882-4.6256 ng/kg) in soil was investigated<br>The addition of white rot fungi following a<br>two-step method can result in a higher degradation<br>rate when compared with the one-step addition<br>strategy</td><td>(Wu et al.,<br>2012)</td></tr><tr><td>Trametes versicolor<br>ATCC 42530</td><td>87 ± 6 (deca-BDE)<br>85 ± 13 (penta-BDE)<br>67 ± 7 (octa-BDE)<br>86% (pre-existent deca-BDE in<br>sewage sludge via biopile<br>systems)</td><td>38h,<br>25°C</td><td>Aqueous phase under<br>aerobic conditions<br>(amber serum bottles)<br>deca-BDE (C<sub>0</sub> = 10<br>mg/L) penta- and octa-<br>BDE (C<sub>4</sub> = 5 mg/L)</td><td>NM</td><td>The addition of Cu<sup>2+</sup> increased the degradation of<br>BDE-209 (69.20%), as copper serves as a cofactor<br>for laccases<br>All congeners were removed at high extents<br>through a mechanism involving the intracellular<br>cytochrome P450<br>In the case of penta- and octa-BDE mixtures,<br>biosorption on the fungus was the major<br>phenomenon resulting in the removal of these<br>pollutants</td><td>(Vilapla et<br>al., 2015)</td></tr></tbody></table>

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<table><caption>Table 2 (continued)</caption><thead><tr><th></th><th></th><th></th><th></th><th></th><th></th><th></th></tr></thead><tbody><tr><td>Phlebia lindneri JN45</td><td>77.3% (20 mg L<sup>-1</sup> of BDE-209)</td><td>30d, 30°C, pH = 4.5</td><td>Liquid Culture Medium</td><td>Hexabromodiphenyl ethers (BDE-154 and BDE-169) &gt; Major products, Lower brominated PBDEs (BDE-47 and BDE-71), OH-PBDEs and open-ring products (2.20, 3.30-tetrahydroxy-4,40-dibromodiphenyl ether, 4-bromopyrogallol, and 3-oxo-4-bromo-adipic acid)</td><td>Biodegradation of deca-BDE (BDE-209, at 0.5-20 mg L<sup>-1</sup>) under the influence of heavy metals and glucose addition was investigated<br>The inclusion of glucose as a co-substrate remarkably enhanced the degradation and removal rates of deca-BDE</td></tr><tr><td>Pycnoporus sanguineus CGMCC 5.00815</td><td>51.26 ± 0.08</td><td>6d, 30°C</td><td>Liquid Culture Medium</td><td>Hydroxylated-PBDEs: 6'-OH-BDE-47<br>2'-OH-BDE-28<br>5'-OH-BDE-47<br>4'-OH-BDE-17<br>Bromophenols: 2,4-DBP, 4-BP</td><td>BDE-209 degradation exhibited a biphasic response to Cu and Pb concentrations, with stimulation at low levels and inhibition at higher concentrations, whereas Cd<sup>2+</sup> inhibits degradation<br>Addition of Cu<sup>2+</sup> (≤5.0 mg L<sup>-1</sup>) enhances laccase activity<br>Ecotoxicity increased at late stages</td></tr><tr><td>Phanerochaete chrysosporium BKM-F-1767</td><td>69.19 (1.0 mg/L of Cd<sup>2+</sup>)</td><td>6d, 30°C</td><td>Tien and Kirk's liquid medium</td><td>Hydroxylated-PBDEs: 6'-OH-BDE-47<br>5'-OH-BDE-47<br>2'-OH-BDE-28<br>4'-OH-BDE-17<br>Bromophenols: 2,4-DBP, 4-BP</td><td>Biodegradation of 2,2',4,4'-tetra-BDE (BDE-47, at 0.5 mg L<sup>-1</sup>) under the influence of heavy metals addition was investigated<br>Addition of Cu++ (1-2 mg L<sup>-1</sup>) enhances laccase activity and promoted protein synthesis and secretion, whereas at C ≥ 5 mg L<sup>-1</sup> BDE-47 removal is inhibited<br>P. sanguineus was able to further remove degradation products: 6'-OH-BDE-47 (22.42%)<br>2'-OH-BDE-28 (23.01%)<br>5'-OH-BDE-47 (27.04%)<br>4'-OH-BDE-17 (27.96%)<br>2,4-DBP (64.21%)<br>4-BP (40.62%)</td></tr><tr><td>Phanerochaete chrysosporium collection number. 5.776</td><td>43.1 ± 0.4% -<br>43.0 ± 1.1% (TCFR+P)</td><td>70d, 30°C</td><td>Tourmaline-catalyzed Fenton-like reaction (TCFR) followed by subsequent mycoremediation in field soil microcosms (TCFR+P) with total C at 18.70 ± 0.08 ng/g</td><td>5 degradation metabolites of BDE-47 and 9 of BDE-153 were identified with C<sub>12</sub>H<sub>20</sub>O as the final product for both pollutants</td><td>Biodegradation of (BDE-47, at 0.5 mg L<sup>-1</sup>) under the influence of cadmium addition was investigated<br>Addition of Cd<sup>2+</sup> enhanced BDE-47 removal when added at low concentrations (≤ 1.0 mg L<sup>-1</sup>), whereas at higher concentrations (≥ 5.0 mg L<sup>-1</sup>) an inhibitory effect was observed<br>Fungal cells revealed higher tolerance to the Toxicity of cadmium compared to extracellular enzymes, as the presence of Cd<sup>2+</sup> significantly impeded the ability of extracellular enzymes (Mn-P and Li-P) to contribute to BDE-47 removal<br>P. chrysosporium was able to further remove degradation products<br>PBDEs with fewer bromine atoms (BDE-28, BDE-47, BDE-100, and BDE-99) were more efficiently removed through TCFR</td></tr><tr><td>Phanerochaete chrysosporium GIM 3.383</td><td>65.65 ± 3.71 (Plant + WRF + BCD<sub>13</sub>)</td><td>65d, 25-32 °C</td><td>Soil Microcosm (Pot experiment)</td><td>NM</td><td>Conversely, PBDEs with a higher degree of bromination were more efficiently removed when a subsequent bioremediation step was used following TCFR<br>For total PBDE removal form soil, TCFR was deemed more efficient than TCFR + P<br>Both approaches resulted in a decrease in soil organic matter, and aromaticity, while surface hydrophilicity, polarity index of humin and humic acid and ergosterol content increased<br>Combined the action of P. chrysosporium with that of a hyperaccumulating plant, Amaranthus hypochondriacus L., in the presence of β-cyclodextrin as a surfactant for the removal of cadmium and deca-BDE (BDE-209)</td></tr><tr><td>Phanerochaete chrysosporium BKM-F-1767</td><td>≥ 90%</td><td>7d, 37°C</td><td>Liquid Medium +/- glucose</td><td>6-OH-BDE-47,<br>5'-OH-BDE-47,<br>4'-OH-BDE-17,<br>2,2',3,3'-OH-BDE-15,<br>2,4-DBP<br>3-oxo-4-bromo-adipic acid</td><td>P. chrysosporium ameliorated the growth of Amaranth, cadmium uptake and BDE-209 degradation<br>β-cyclodextrin increased soil Mn-P activity<br>A positive correlation between BDE-209 degradation and β-cyclodextrin and Mn-P activity was observed<br>The impact of rhamnolipids as bioactiveants on the removal of BDE-47 (at 100 µg) using P. chrysosporium as a remediation agent was investigated<br>At low concentrations (0.05-0.1 CMC) (critical micelle concentration), rhamnolipids enhanced the removal of BDE-47, whereas when introduced at higher concentrations (0.5-1 CMC) rhamnolipids inhibited the removal of BDE-47, both with and without the addition of glucose<br>At low concentrations, rhamnolipids promote LMEs activities (Mn-P, Li-P, and Lac)</td></tr></tbody></table>

Abbreviations. 1,2,4-trihydroxybenzene (1,2,4-THB), 2,7-dichlorodibenzo-p-dioxin (2,7-diCDD), 2,8-dichlorodibenzofuran (2,8-diCDF), dibenzo-p-dioxin (DD), chlorobenzyl alcohols (CB-OHs) and chlorobenzaldehydes (CB-CHOs), 2,4-dibromophenol (2,4-DBP), 4-bromophenol (4-BP).

retardants (BFRs) compounds, widely used in electronic products, textiles, plastics, and fabrics to prevent ignition and damages caused by fire (Vilaplana et al., 2013). PBDEs directly mix in rather than chemically bond to commercial/industrial products, therefore they tend to easily migrate from such products to environmental matrices following their disposal (Yao et al., 2021). PBDEs' solubility decreases with the number of bromine substitutions, whereas their toxicity increases with the latter. Among PBDEs, penta-, octa-, and deca-BDEs mixtures are pre-eminently detected in environmental samples, thus their recognition as persistent organic pollutants by
<figure><img src="image_4.png" /><figcaption>**Fig. 2.** Phylogenetic distribution of fungal species reviewed for persistent organic pollutant degradation. Brown colored bars represent the relative abundance of publications for each fungal species presented in this tree (Level 1). Colored ranges (blue, yellow, peach) identify fungal phyla within the tree (Level 2). Colored strips (brown, dark blue, clear yellow) along branches indicate the type of POP the fungi are capable of degrading (agricultural, industrial, or by-product) (Level 3). The phylogenetic tree was generated with the maximum-likelihood method by FastTree and uploaded to iTOL (Interactive Tree of Life).</figcaption></figure>
the Stockholm Convention.

Fungal degradation of PBDEs has been primarily reported in relation to the fungal species *Trametes versicolor*. The first evidence of PBDEs degradation by WRF was reported by Hundt *et al.* (1999) who observed the ability of *T. versicolor* to degrade 4-bromo-BDE through hydroxylation reactions mediated by the intracellular enzyme cytochrome P450 monooxygenase, followed by ring cleave to yield 4-bromophenol as the final degradation product. Afterward, several studies have explored the potential of fungal remediation strategies for PBDE degradation. More detailed information on the degradation of this pollutant is provided in Table 2 (section colour code: orange), where we have listed the fungal species involved or shown potential in its biodegradation. This section also presents the percentage of degradation, the incubation conditions (time and temperature), the culture medium or culture settings used, the degradation products identified (when known), as well as additional notes.

# 5. Knowledge gaps and new strategies for identifying promising fungal degraders of persistent organic pollutants

## 5.1. Potential of fungal bioremediation for emerging POPs

This review comprehensively tackles the current understanding of mycoremediation for POPs, with special emphasis on those impacting soil health, either through direct application on soils to enhance agricultural productivity (agricultural POPs in 3, Fig. 1 and Table 1), or through on-site release in operational industrial plants, orphaned sites and landfill soils, or through wet and dry air depositions (industrial or unintentional POPs in 4 and Table 2). Thus, we delve into 10 of the 12 initial POPs listed under the Stockholm Convention and given the continuously evolving landscape of POPs, eight of the newly added 22 POPs (as of May 2023) for which fungal degradation capacities has been extensively reported were also addressed whether solely or as part of a larger family of

<figure><img src="image_5.png" /></figure>

**Fig. 3.** Approaches allowing isolation and characterization of fungal strains of interest in soil bioremediation. Fig. 3 | Description | Following the sampling campaign (a), the isolation and culture of fungal isolates can be achieved whether through traditional cultures or the culturomics approach (b). The subsequent identification of grown colonies can be accomplished either by biochemical assays among which MALDI-TOF (c) has emerged as a rapid powerful identification tool or through amplification and sequencing of target barcode genes such as ITS, β-tub, TEF, or D1/D2 among others (d). Identified isolates (e) can be thus selected based on their ability to tolerate high concentrations of a given pollutant. (f) Tolerance tests compare the radial extension rate and the macroscopic morphology of fungal isolates grown in the presence of increasing concentrations of the pollutant to those of a control grown without the pollutant. Tolerant isolates are then subjected to degradation assays (g) where the ability of the isolates to degrade the pollutant is tested. (h) Among tested isolates, those displaying high rates of degradation can be further screened for their ability to produce LMEs through qualitative assays based on color change reactions specific for each lignin modifying enzymes, or through quantitative assays measured spectrophotometrically (not shown). (i) The uncovering of the degradation pathways underlying the degradation process can be done through various approaches, among which the use of inhibitors of the intracellular pathway (1-aminobenzotriazole, piperonyl butoxide) or stimulators of the extracellular pathway (low-nitrogen content) during degradation assays, allows to evaluate the contribution of each pathway in the degradation extent of the pollutant. Finally, microcosm experiments (j) can be carried out in order to further characterize the remediation ability of isolates, as microcosm trials offer an intermediary between degradation assays performed in liquid culture media and the application of remediation on large-scale fields.
compounds.

Given their extended half-lives, their hydrophobicity and chemical stability, POPs bind to soil organic matter, severely impacting soil microbial communities, nutrient cycles and soil physico-chemical properties and overall soil health. Fungal-based bioremediation strategies are increasingly being explored as sustainable solutions to decontaminate soils impacted by POPs and to restore soil biodiversity and functionality as part of the One Health approach.

Fig. 2 provides a comprehensive overview of fungal species exhibiting the capabilities to degrade these POPs, according to level 1 (number of studies), level 2 (fungal phyllum) and level 3 (class of POPs) discussed within the scope of this review. It highlights that basidiomycota (i.e. *trametes versicolor*, *phlebia lindtneri*, *phanerochaete chrysosporium*, *pleurotus ostreatus*) were the most applied fungal phyllum, while ascomycota (mostly penicillium) were less tested. Furthermore, while a wide variety of strains have been tested for agricultural POPs, the spectrum is much more limited for industrial and unintentional POPs. For future work, this overview can help to select potential strains for a given pollutant or versatile strains for a mixture of pollutants, as is generally the case in-situ, or to extend knowledge of the degradation efficiency of a given strain by testing other pollutants.

While this review strives to provide a comprehensive resource on mycoremediation of the POPs listed under the Stockholm Convention, some historical or emerging POPs, either newly listed under the Stockholm Convention or currently candidate POPs were omitted because of a lack of studies. However, current research progress for these emerging POPs is discussed in this section where data is available.

In the case of historical POPs, Mirex and Toxaphene, listed since 2001 as part of the dirty dozen under the Stockholm Convention, Kennedy et al. (1990) and Romero et al. (2006) revealed that *Phanerochaete chrysosporium* and *Bjerkandera sp.* strain BOL13 displayed the ability to degrade Mirex and Toxaphene, respectively. The WRF *Phlebia lindtneri* has been reported as capable of degrading polychlorinated naphthalenes through hydroxylation reactions (Mori et al., 2009).

Concerning, the newly added persistent organic pollutants such as Dechlorane Plus, Dicofol, Hexabromobiphenyl, Hexabromocyclododecane, Hexachlorobutadiene, Methoxychlor, Pentachlorobenzene, Per- and Polyfluoroalkyl Substances, Polychlorinated naphthalenes, and Short-chain chlorinated paraffins, lack extensive and in-depth literature on fungal transformation due to their recent inclusion. While a comprehensive and in-depth picture remains elusive, several studies have reported the existence of fungal isolates capable of degrading some of these POPs.

For Dicofol, a newly listed POP, *Trametes versicolor* (Hu et al., 2020), in addition to cellulase extracts from *Trichoderma longibrachiatum* (Wang et al., 2015) have been demonstrated to display degradative capacities. In the case of Methoxychlor, an organochlorine insecticide introduced most recently at the eleventh meeting of the Stockholm Convention (2023), the fungal species *Stereum hirsutum* (Lee et al., 2006), in addition to *Phanerochaete chrysosporium* (Hirai et al., 2004; Grifoll and Hammel, 1997) have been reported to display degradation capacities towards this newly listed POP. Regarding hexabromocyclododecane, *Ganoderma lucidum* and *Antrodia cinnamomea* have been reported as capable of adsorbing this POP rather than transforming it (Chou et al., 2021). The aryl chloride pentachlorobenzene has been shown to degrade through the action of the WRF *Trametes versicolor* and its arsenal of LMEs (Sari et al., 2016).

Per- and polyfluoroalkyl substances (PFAS), used in a variety of products such as fire-fighting foams, Teflon, and food packaging, PFAS, also known as forever chemicals, are notoriously resistant to degradation. This group of chemicals includes Perfluorohexane sulfonic acid (PFHxS), its salts and PFHxS-related compounds, Perfluorooctane sulfonic acid (PFOS), its salts and perfluorooctane sulfonyl fluoride (PFOSF), along with Perfluorooctanoic acid (PFOA), its salts and PFOA-related compounds, it has been shown that fungal transformation of these newly listed POPs involves enzyme-catalyzed oxidative humification reactions (ECOHR), which consist of a series of oxidative reactions carried out by the arsenal of extracellular enzymes (LMEs), namely Lac, Mn-P, and Li-p, typically secreted by WRF and BRF as these LMEs convert chemical compounds found in soil systems to reactive radicals which eventually react with PFASs (Grgas et al., 2023). In this context, fungal species such as *Aspergillus niger* (Tseng, 2012) and *Phanerochaete chrysosporium* (Tseng, 2012; Zhou et al., 2023) have been reported as species with potential degradative capacities against this group of POPs. A recent study indicated that certain white rot fungi, namely *Irpex lacteus*, can degrade these compounds in the presence of plant-based carrier materials (Li et al., 2022). This same study demonstrated that chemically modified lignocellulose can adsorb PFAS, serving as a support for a fungus capable of degrading them. After two weeks, more than 98 % of PFOA and 99 % of PFOS were eliminated, suggesting effective biological detoxification and the transformation of PFAS into less toxic products.

Emerging POPs, like long-chain perfluorocarboxylic acids (LC-PFCAs), Medium Chain Chlorinated Paraffins (MCCPs), and Chlorpyrifos are currently candidate POPs under review and therefore haven't been officially listed under the Stockholm Convention. In the case of the latter, its current state as candidate POP hasn't prevented a significant amount of research from exploring the potential of fungal degradation of this pesticide whose application dates back to 1965 (U.S. Environmental Protection Agency [EPA], 2023). Fungal strains such as *Aspergillus sp.* F1, *Penicillium sp.* F2, *Penicillium sp.* F3, *Eurotium sp.* F4, and *Emericella sp.* F5 (Maya et al., 2012), *Cladosporium cladosporioides* Hu-01 (Chen et al., 2012), *Acremonium sp.* GFRC-1 (Kulshrestha and Kumari, 2011), *Trichosporon spp.* (Xu et al., 2007), *Verticillium* sp. DSP (Yu et al., 2006), *Byssochlamys spectabilis*, *Aspergillus fumigatus* (Kumar et al., 2021), *Aspergillus sydowii* (Soares et al., 2021), *Lenzites betulinus*, *Irpex lacteus*, and *Phlebia sp.* (Wang et al., 2020) have been reported to degrade Chlorpyrifos to different extents under different conditions.

Besides emerging POPs, microplastics (MPs) are tiny plastic particles, typically smaller than 5 mm, that are intentionally added to certain product or produced from the degradation of plastic waste (Xiang et al., 2023). MPs have POP-like properties, being highly bioaccumulable and toxic for marine and terrestrial ecosystems, their low or long-term degradability and their capacity to be transported over large geographic distances by waters and air (Siwach et al., 2024). Moreover, MPs can interact with POPs by sorption equilibrium or even contain additives, that are listed as POPs, such as the brominated flame retardants (Bouhoum et al., 2019; Gateuille and Naffrechoux, 2022). It is now widely reported that MPs are disseminated in the whole environment, including in
particular agricultural and landfill soils (Athulya et al., 2024). Plastics are difficult to biodegrade, but some fungal species, such as *Aspergillus* and *Penicillium*, have been shown to degrade certain polymers and plastic additives. For example, *Aspergillus flavus* PEDX3, isolated from the digestive tract of the wax moth (*Galleria mellonella*), has been shown to degrade high-density polyethylene (HDPE) microplastics (Zhang et al., 2020). In addition, several other fungal species belonging to the genera *Aspergillus*, *Penicillium*, *Bjerkandera*, *Cladosporium*, *Phanerochaete*, *Pleurotus*, and *Agaricus* have been reported to be effective in degrading various types of plastics (Srikanth et al., 2022).

## 5.2. Insights on novel approaches allowing isolation and characterization of fungal strains of interest in bioremediation

The knowledge gaps and focus on emerging POPs highlight the ongoing challenge of keeping current with the evolving POP landscape as well as the crucial need for effective mycoremediation strategies. In this context, isolation of fungal strains with efficient and targeted remediation abilities remains a critical challenge. Therefore, this paragraph will explore insights and innovative approaches for isolating and characterizing these potent bioremediation tools.

Most studies on mycoremediation focused on white rot fungi because of their ability to secrete nonspecific ligninolytic enzymes capable of oxidizing various aromatic compounds (Čvančarová et al. 2012; Sadaňoski et al., 2019; Benitez et al., 2021). However, their effectiveness in real conditions remains limited due to their low competitiveness against native microbiota and their sensitivity to unfavorable environmental conditions. (Mouhamadou et al., 2013; Sage et al., 2014). Therefore, recent studies have focused on the use of indigenous strains belonging to the Ascomycota and Zygomycota divisions, which are better adapted to contaminated environments (Germain et al., 2021b; Périgon et al., 2019; de la Cruz-Izquierdo et al., 2021). Thus, the first challenge in the search for fungal strains of interest for soil bioremediation is their isolation from environmental matrices. The isolation of stains from soils contaminated with POPs has a considerable advantage, as these isolates have already been subjected to natural selection that favours their adaptation to the pollutants present. After sampling of natural soils (Fig.3.a), several isolation methods are used, based on the use of specific culture media (Fig. 3b). Among those commonly used are Malt Extract Agar (Germain et al. 2021b; Périgon et al., 2019; Sage et al., 2014; Mouhamadou et al., 2013), Potato Dextrose Agar (Aranda et al., 2017; Qiu et al., 2021), Minimum Medium supplemented with the pollutant as selective agent (de la Cruz-Izquierdo et al., 2021; Lahkar and Deka, 2016; Wu et al., 2009), Czapeck-Dox Medium (Mineki et al., 2014; Sagar and Singh, 2011; Goswami et al., 2009), Sabouraud Dextrose Agar (Agrawal and Shahi, 2017), and Martin Agar Medium (Kataoka et al., 2010; Borowik et al., 2017).

The subsequent identification of grown fungal colonies calls for the use of biochemical assays (Fig. 3c) (Borowik et al., 2017; Lima et al., 2019) or barcode gene amplification and sequencing (ITS, β-tubulins, TEF, D1/D2) (Fig. 3d) (de la Cruz-Izquierdo et al., 2021). Among biochemical assays, matrix-assisted laser desorption ionization-time of flight mass spectrometry (MALDI-TOF MS) (Fig. 3c) has emerged as a rapid and potent tool for microbial identification and typing, revolutionizing clinical mycology and presenting a potential in fungal ecology studies (Cassagne et al., 2011).

Afterward, identified colonies undergo a series of tests and assays allowing for the selection among isolates, those harboring the ability to transform and degrade pollutants at high rates (Fig. 3f). Tolerance tests can potentially serve as a starting point for the selection of isolates based on their ability to grow in the presence of increasing concentrations of a given pollutant (de la Cruz-Izquierdo et al., 2021).

Tolerant isolates are then selected to undergo degradation assays (Fig. 3g), where isolates are tested for their ability to degrade and transform pollutants in liquid medium, and pollutant removal rates are measured using chromatographic techniques (gas or liquid chromatography coupled to simple or tandem mass spectrometry) (Wang et al., 2019; de la Cruz-Izquierdo et al., 2021; Agrawal and Shahi, 2017). To avoid bias in the determination of the degradation rate, abiotic and heat killer controls must be used to distinguish adsorption from active metabolism of pollutants (J. Germain et al., 2021b).

To insure active metabolism of the strains, the ability to secrete the extracellular arsenal of lignin-modifying enzymes is considered an interesting feature for any fungal isolate in the context of bioremediation (Fig. 3h). The enzymatic activity of the isolates can be assessed using qualitative and quantitative ligninolytic enzyme (LME) tests, involving color change reactions or spectrophotometric measurement of specific enzyme activities, respectively. In the case of the former the apparition or not of a colored zone around the fungal diameter disc along with its diameter, inform about the ability or not of the fungus to produce a given LME along with the intensity of production. On the other hand, the quantitative estimation, which provides a more accurate assessment of the ligninolytic

<table><caption>**Table 3**
Comparison of Bioremediation Capabilities between Bacteria and Fungi.</caption><thead><tr><td>Criteria</td><td>Bacterial Bioremediation</td><td>Fungal Bioremediation</td><td>Reference</td></tr></thead><tbody><tr><td>Types of enzymes</td><td>Hydrolases, specific oxidoreductases</td><td>Ligninolytic enzymes (laccases, peroxidases) and other intracellular (P450 monooxygenases) and extracellular enzymes</td><td>(Kim and Pascual, 2021; Mousavi et al., 2021)</td></tr><tr><td>Environmental adaptability</td><td>Sensitive to extreme variations</td><td>More resistant to acidic/saline conditions and toxins</td><td>(Branco et al., 2022; De Doctorat, 2018)</td></tr><tr><td>Degradation of complex substrates</td><td>Less effective on recalcitrant structures</td><td>Capable of degrading complex structures such as polycyclic aromatic hydrocarbons</td><td>(Cerniglia and Sutherland, 2010; Mousavi et al., 2021)</td></tr><tr><td>Degradation rate</td><td>Usually fast</td><td>Often slower but more complete</td><td>(Dell'anno et al., 2022; Mousavi et al., 2021)</td></tr><tr><td>Industrial applications</td><td>Effluent treatment, biodegradation of solvents</td><td>Bioremediation of soils, contaminated waters and organic waste</td><td>(Mousavi et al., 2021)</td></tr></tbody></table>
activity, consists of measuring spectrophotometrically following incubation in the presence of adequate enzymatic substrates and required supplements (Agrawal and Shahi, 2017; Germain et al., 2021b).

To better understand the degradation mechanisms, several studies have evaluated the contribution of extracellular and intracellular pathways to the transformation of pollutants (Fig. 3i). For instance, de la Cruz-Izquierdo et al. (2021) tested the involvement of intracellular and extracellular enzymes in the degradation of phenanthrene by *Aspergillus flavipes* and A. *oryzae*. Other approaches explored the effect of limited nitrogen on extracellular enzyme induction and the impact of CYP450 monooxygenase inhibitors on the intracellular pathway (Xiao and Kondo 2020b). In another strategy, Wang et al. (2019) evaluated the intracellular and extracellular distribution of pollutants in cultures of *Pycnoporus sanguineus*, complementing this work with chromatographic analyses (GC-MS) to identify intermediate metabolites.

In order to assess the residual toxicity after biotransformation of pollutants, ecotoxicity tests can be carried out with sensitive organisms such as *Ganoderma lucidum* or *Vibrio fischeri*, allowing the toxicity of metabolic products to be compared with that of the initial pollutant (Čvancarová et al. 2012; Sadanoski et al., 2019).

Finally, microcosm approaches represent an important step in the evaluation of the in-situ bioremediation capacity of a given fungal species (Fig. 3j). In fact, microcosm studies constitute an intermediary step preceding large-scale application, thus their importance for the verification of the remediation ability of tested isolates in in-situ conditions rather than culture media. These approaches allow for a better understanding of the processes involved and the selection of the most promising strains.

### 5.3. Mycoremediation beyond the lab: field applications and comparative insights with bacterial bioremediation

Although most studies on fungal bioremediation are conducted in the laboratory, several field applications have demonstrated its potential. For instance, fungi have been applied for the decontamination of soils contaminated by hydrocarbons. Gregorio et al. (2016) explored the use of spent mushroom substrates from industrial production of *Pleurotus ostreatus* as a structuring agent in a dynamic biopile. This biopile treated 7 tons of soil historically contaminated with PAHs (6469 ± 423 mg PAH per kg), classified as hazardous waste. After 8 months, PAH concentrations were significantly reduced to 112 ± 5 mg PAH per kg, making the soil suitable for rein-troduction to the original industrial site. This approach highlights the effectiveness of using spent mushroom substrates for the decontamination of hydrocarbon-polluted soils (Gregorio et al., 2016). Similarly, a field study by Hestbjerg et al. (2003) evaluated the efficacy of the same species *P. ostreatus* for the degradation of polycyclic aromatic hydrocarbons (PAHs) in contaminated soils under real-world conditions. The fungus was introduced as commercial mushroom production waste, mixed with the contaminated soils. After nine weeks, the concentrations of 3-, 4-, and 5-6-ring PAHs were reduced by 78%, 41%, and 4%, respectively, in one of the *P. ostreatus*-treated sites, with notable correlations with elevated laccase enzyme activity and soil acidification. These reductions were also accompanied by significant increases in phenanthrene- and pyrene-degrading bacterial populations, suggesting that the fungus may promote endogenous microbial activity involved in bioremediation (Hestbjerg et al., 2003).

Fungal remediation has been also applied on the field-scale for industrial effluent treatment. For instance, the use of the WRF *Trametes versicolor* has been explored for the treatment of industrial effluents, particularly for the degradation of textile dyes and pharmaceutical residues. Studies have shown that this fungus has a remarkable ability to decolorize various synthetic dyes used in the textile industry (Amaral et al., 2004; Llorens-Blanch et al., 2015). Amaral et al. (2004) evaluated the efficiency of *T. versicolor* in decolorizing three synthetic textile dyes in the presence and absence of glucose. The results showed a decolorization of approximately 97%, indicating that fungal metabolism, stimulated by the presence of glucose, plays a crucial role in this process. Furthermore, the application of *T. versicolor* to real wastewater from the textile industry led to a decolorization of 92%. These results highlight the potential of *T. versicolor* for the effective treatment of textile effluents, thus contributing to the reduction of environmental pollution (Amaral et al., 2004). Moreover, biofiltration systems combining bacteria and fungi have been investigated for the treatment of volatile organic compounds (VOCs) in industrial effluents. Wang and Xue (2024) compared the performance of bacterial biofilters alone and coupled bacteria-fungi biofilters for the removal of gas mixtures containing benzene, toluene, ethylbenzene and para-xylene (BTEp-X). The results showed that the coupled biofilters exhibited higher removal efficiency and increased operational stability over a six-month period, indicating their potential for large-scale industrial applications.

Fungal bioreactors have also been used for water treatment. In fact, the use of fungal biofilters has shown promising results for the removal of heavy metals and organic pollutants in wastewater. These fungal biofilters exploit extracellular enzymes produced by fungi to oxidize and decompose a wide range of persistent organic compounds. Hu et al. (2020) demonstrated the ability of *T. versicolor* to remove 83% of diuron and 100% of its main metabolite, 3,4-dichloroaniline, after a 7-day incubation in a trickling bed reactor operating under non-sterile conditions, representing an advance in the field of bioremediation techniques for pesticide-contaminated water using fungal reactors on a real scale (Hu, Toran et al., 2020).

Bioremediation relies on the ability of microorganisms to transform or eliminate pollutants. In this context, bacteria and fungi exhibit notable differences (Table 3). Fungi, especially basidiomycetes and ascomycetes, offer a particular advantage in the degradation of complex chemical structures thanks to their extracellular enzymes. However, bacteria are often faster and easier to cultivate under industrial conditions. A combined approach could offer optimal solutions depending on the type of pollution.

## 6. Conclusions

Owing to their extreme metabolic versatility, their unique biochemical features as well as their ecological advantage over bacteria, much attention has been paid to the application of fungi as potential bioremediation agents over the past few decades. The selection of the ideal fungal strain for soil remediation is based on several criteria: (i) its ability to tolerate and withstand the presence of the
contaminant; (ii) its non-pathogenicity; (iii) its ability to colonize and grow abundantly in the telluric environment; (iv) its ability to reduce the pollutant's associated toxicity; (v) and finally its ability to degrade the pollutant at exceedingly high rates.

The choice of the initial pollutant concentration proves is a crucial step in assessing the bioremediation potential of microorganisms. The adoption of standardized concentrations as well as the establishment of minimum thresholds specific to each class of pollutants, are essential to allow a reliable comparison of degradation capacities between different strains. In the absence of such standards, the variability of experimental protocols complicates the interpretation of results and makes any attempt at comparison almost impossible. Indeed, the high degradation rates reported in some studies may be biased by the use of low initial concentrations rather than by the intrinsic capacity of the strains to effectively degrade the pollutant. This bias can lead to an overestimation of the bioremediation potential of some isolates and distort the assessment of their effectiveness in real conditions. Furthermore, although liquid culture tests are a useful preliminary step to screen the biodegradation capacities of fungal isolates, these studies should systematically be supplemented by microcosm experiments. These would allow to evaluate the capacity of the selected strains to colonize natural matrices, to adapt to environmental conditions and to compete with indigenous microbial communities. The integration of these standardized criteria would considerably improve the robustness of mycoremediation studies and would promote the development of more reproducible and transferable approaches on a large scale. While only a handful of fungal species, primarily belonging to the WRF group, have been extensively explored for POP remediation, further research is needed to leverage the bioremediation potential of a wider fungal diversity. Recent advancements in fungal isolation techniques, through High-throughput culture approaches, such as culturomics, combined with rapid identification techniques like MALDI-TOF MS, hold immense potential for expanding the soil fungal repertoire and paves the way for isolating novel strains with superior remediation capabilities.

### CRediT authorship contribution statement

**Malleret Laure:** Writing – review & editing, Validation, Supervision, Project administration, Methodology, Funding acquisition, Conceptualization.
**Elias Charbel:** Writing – original draft, Visualization, Methodology, Investigation, Formal analysis, Data curation.

**Ranque Stéphane:** Writing – review & editing, Supervision, Resources, Project administration, Funding acquisition, Conceptualization.

### Funding

This research was funded by the Institut Hospitalo-Universitaire (IHU) Méditerranée Infection and by the French Government under the "Investissements d'avenir" (Investments for the Future) programme managed by the Agence Nationale de la Recherche (ANR, fr: National Agency for Research), (reference: Méditerranée Infection 10-IAHU-03) and the Région Provence-Alpes-Côte d'Azur, European ERDF funding (European regional development fund) and PRIMMI (Plateformes de Recherche et d'Innovation Mutualisées Méditerranée Infection).

### Declaration of Competing Interest

The authors declare that they have no known competing financial interests or personal relationships that could have appeared to influence the work reported in this paper.

### Acknowledgement

EC is a recipient of a PhD scholarship from the Fondation Méditerranée Infection.

### Appendix A. Supporting information

Supplementary data associated with this article can be found in the online version at doi:10.1016/j.eti.2025.104247.

### Data availability

Data available in tables and SI.

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