Bioremediation of Polycyclic Aromatic Hydrocarbons (PAHs) in Aqueous Environments: A Review of Biofiltration, Biosorption, and Biodegradation Strategies Using Living Fungal Mycelium

Abstract

Accelerating urbanisation and industrial activity have led to the widespread release of polycyclic aromatic hydrocarbons (PAHs), a class of persistent organic pollutants with serious ecological and health consequences. While physical and chemical remediation techniques are widely used, they often require nonrenewable resources and generate secondary waste. Fungal-based bioremediation offers a promising alternative, leveraging the unique metabolic pathways and structural properties of fungi to break down or adsorb PAHs. This review focuses on three strategies of PAH remediation in aquatic environments: biofiltration, biosorption, and metabolic degradation. We conduct a comparison between conventional systems and fungal approaches with reference to the literature (2000–2025). Fungal matrices are identified as being able to capture and adsorb PAHs, facilitating localised remediation that capitalises on the biological capabilities of fungal organisms while requiring lower resource inputs than conventional methods. This review highlights fungal matrices as multifunctional water filtration membranes and provides insights for the application and development of engineered living materials (ELMs) for the water detoxification of PAHs.

1. Introduction

1.1. Biological Strategies for Water Remediation of PAHs: Biofiltration, Biosorption, and Biodegradation

PAHs are persistent, hydrophobic organic pollutants with carcinogenic, mutagenic, and teratogenic properties. Their presence in surface water, sediments, and groundwater poses significant ecological and public health risks [1,2]. Due to their low water solubility and high molecular stability, PAHs tend to adsorb onto particles and sediments, making them resistant to conventional remediation methods such as chemical oxidation, photodegradation, or physical filtration [3]. Traditional remediation technologies, such as microfiltration, adsorption, or chemical treatments, often rely on non-renewable inputs and can generate toxic by-products or cause environmental disturbance [4,5]. Bioremediation, by contrast, is increasingly being recognized as a sustainable and effective approach for degrading and removing PAHs from soil and water. It leverages the metabolic activity of microorganisms, including bacteria, algae, and fungi, to transform hazardous compounds into less harmful products [6,7]. Among biological agents, fungi offer distinct advantages in PAH remediation due to their ability to produce extracellular oxidative enzymes (e.g., laccases, manganese peroxidases), form extensive mycelial networks, tolerate toxic environments, and biosorb organic compounds through their cell walls [8]. Fungal-based PAH bioremediation offers a complementary strategy to conventional treatments [9], and the key strategies include biofiltration, biosorption, and biodegradation, each exploiting the unique physiological and biochemical properties of fungi; however, their use is limited due to the fact that its effectiveness is related to contaminant bioavailability, environmental conditions, and logistics (Figure 1). In biofiltration, biological mass, often integrated into a fixed support or membrane, acts as a filter, physically trapping PAHs while facilitating their interaction with active sites or microbial consortia [10]. Biosorption involves the passive binding of PAH molecules to fungal cell walls, especially to functional groups, such as carboxyl, hydroxyl, and amine groups, found in chitin and glucans [11]. This process does not require metabolic activity and is effective even under non-ideal environmental conditions that would inhibit biological growth and enzymatic activity. Biodegradation, by contrast, depends on enzymatic transformation of PAHs into less toxic compounds. Fungi, particularly white-rot species, secrete oxidative enzymes, such as laccases, lignin peroxidases, and manganese peroxidases, that break the complex aromatic structures of PAHs [12]. Together, these strategies enable fungi to act not just as passive adsorbents but as active participants in pollutant transformation. Their potential to function in acidic, saline, or low-nutrient environments makes them particularly suited for in situ or decentralised water treatment systems. Fungal matrices are identified as being able to capture, adsorb, and enzymatically transform PAHs, thereby enabling multifunctional localised remediation with lower resource inputs than conventional methods. This paper reviews the literature (2000–2025) on biological remediation strategies for PAHs and places special focus on unfolding relevant strategies for using living fungal matrices as active filtration membranes in aqueous environments.

Studies were retrieved from Web of Science, Google Scholar, and PubMed using search terms including “fungi”, “PAH”, “bioremediation”, “biodegradation”, “biosorption”, and “biofiltration” (Figure 2). Through the analysis of biofiltration, biosorption, and enzymatic degradation, we highlight their mechanisms, efficiencies, and applications. The review addresses fungal remediation strategies, emphasising the prospective integration of these three mechanisms into one process utilising living fungal mycelium membrane. It aims to evaluate the plausibility of combining these mechanisms into an integrated water system of fungi-based living membranes, providing a valuable resource for advancing engineered living materials (ELMs) for water decontamination and to guide the future design of biologically active PAH remediation systems.

1.2. PAHs

PAHs are hydrophobic organic contaminants consisting of multiple fused benzene rings. They are generated mainly by the incomplete combustion of organic matter, including fossil fuels, wood, and biomass, and are released through industrial processes, transportation, agriculture, and domestic activities [2,13] (Figure 3). Though they can also originate from natural events (e.g., wildfires or volcanic activity), the anthropogenic contribution to PAH pollution has increased significantly since the industrial era, with environmental concentrations now estimated to be 3–20 times higher than pre-industrial levels [14]. In the atmosphere, PAHs occur in both gaseous and particle-bound forms; low-molecular-weight (LMW) PAHs remain mostly in the gas phase, while high-molecular-weight (HMW) PAHs preferentially adsorb to particulates. After deposition, their hydrophobicity and strong affinity for organic carbon lead to accumulation in soils and sediments. In aquatic systems, PAHs enter via atmospheric deposition, oil spills, and urban or industrial runoff. LMW PAHs, being more soluble, partition into the water, whereas HMW PAHs associate with particles and settle into sediments, creating long-term sinks that are difficult to remediate and prone to bioaccumulation. These organic pollutants are listed as priority by environmental agencies worldwide, and 16 are spread extensively [15] (Figure 4). The physicochemical properties of PAHs include their molecular weight, structure, and bonding patterns, which play a critical role in their environmental behaviour and toxicity. LMW PAHs (two–three rings) are more volatile and water-soluble, while HMW PAHs (four or more rings) are more stable, less soluble, and often more carcinogenic [2]. Structurally, PAHs can be categorized by ring arrangement into linear (e.g., anthracene), clustered (e.g., pyrene), or angular forms (e.g., dibenz[a,h]anthracene). Angular PAHs tend to be thermodynamically stable, yet possess reactive bay regions, concave pockets at ring junctions that offer points of attack for fungal oxidative enzymes [16]. This structural insight is key to understanding why fungal biodegradation strategies are effective, as enzymes like laccases target these vulnerable sites to initiate degradation. Their effective removal from aquatic environments remains a critical challenge, one for which fungal bioremediation offers a promising complementary solution.

2. Biofiltration and Other Methods

2.1. Overview of Biofiltration

Traditional filtration methods rely primarily on physical or chemical processes, and they often involve intricate procedures, substantial costs, and high energy consumption (

Table 1. Physical filtration methods, their energy efficiency limitations, and references.

Physical Filtration	Energy Efficiency	Methods	Limitations	References
Membrane filtration	Energy intensive in reverse osmosis (RO); moderate for ultrafiltration/microfiltration	Semi-permeable membranes to separate contaminants	High operational costs due to membrane fouling and frequent replacement	Baker, 2023 [25]
Gravel and multimedia filtration	Energy efficient	Phisical entrapment of through the porous bed,	Ineffective towards ions or molecules dissolved in water, the lifespan is dictated by physical clogging	Hatt et al., 2007 [26] M. Lazim, 2020 [27]
Slow and rapid sand filtration	Energy efficient	Water passes through layers of gravel, sand, and anthracite	Limited ability to remove very fine particles and dissolved substances	Kawamura, 2000 [28]
Activated carbon filtration	Energy efficient in use, energy intensive in production	Adsorbs contaminants onto the porous surface of activated carbon	High cost, limited lifespan; requires frequent regeneration or replacement	Tarikuzzaman, 2023 [29]

and

Table 2. Chemical filtration methods, their energy efficiency, limitations, and references.

Chemical Filtration	Energy Efficiency	Methods	Limitations	References
Coagulation and flocculation followed by filtration	Moderately energy-intensive	Coagulants like alum or ferric chloride neutralize particle charges, forming larger aggregates (flocs) that are filtered out	Chemical costs and sludge disposal issues; requires precise dosage	Gregory & Duan, 2003 [30]
Ion exchange filtration	System related. In dilute streams there is a higher energy consumption per unit ion removed	Ion-exchange membrane fundamentals	Concentration polarisation, low current efficiency with complex ions, scaling/fouling, membrane heterogeneity	Pismenskaya & Nikonenko, 2021 [31]
Chemical precipitation	Moderately energy-intensive	Contaminants are converted into insoluble precipitates through chemical reactions and then filtered or settled out	High sludge production; requires proper chemical handling	Wang, 2005 [32]
Oxidation	Moderately/high energy-intensive	Advanced oxidation processes (AOPs) using engineered nanomaterials (ENMs) as catalysts. Risk of forming halogenated or toxic oxidation products	Might require large surface area	Pismenskaya & Nikonenko, 2021 [31]

). Biofiltration, by contrast, is gaining attention for its numerous advantages, as this method relies on biological entities and eliminates the need for chemical additives while being both energy efficient and effective in contaminant removal [17]. At its core, biofiltration is a physical process that relies on the activity of diverse microbial communities that colonise the filter media to break down pollutants [18]. The efficiency of this process is influenced by several factors, including the composition, activity and resilience of the microbial community. Typically, these communities are primarily dominated by bacteria and archaea, while fungi and other eukaryotic organisms play secondary but vital roles [19]. Biofiltration not only addresses water contamination challenges but also aligns with sustainable and environmentally friendly practices in water treatment such as membrane bioreactors, anaerobic digestion, and advanced oxidation processes [20]. Different factors influence the efficiency of the filter and its microbial community. The availability of nutrients such as nitrogen, phosphorus, and vitamins in the water is crucial for sustaining the community’s metabolism. Temperature plays a significant role, as low temperatures can inhibit the metabolic rate of microbiomes. Optimal pH ensures the metabolic activity related to the type of the microbial community; in fact, extreme pH can potentially inhibit growth. Additionally, the filter media’s material composition and its physical and chemical properties shape the microbial environment. The media’s characteristics, including porosity, surface area, and roughness, affect how microbiomes attach to the filter and determine which microbial communities will thrive [21,22,23]. Lu et al. (2025) systematically demonstrated that the liquid-holding capacity (LHC) and BET surface area of lignocellulosic carriers are quantitative predictors of biofilm productivity, with higher values directly enhancing microbial attachment and metabolic activity. This provides measurable criteria for selecting carriers and links material choice to predictable remediation performance [24].

The microbiome of water biofilters is shaped by environmental factors and engineering design choices. Biofiltration technologies vary in design and application, each suited to specific filtration needs, and host distinct microbial communities that contribute to their effectiveness. One widely used technology is Rapid Sand Filtration (RSF), which operates at high flow rates of 5 to 30 m per hour. RSF efficiently separates harmful organic and inorganic particles and is dominated by microbial communities primarily composed of Proteobacteria and Nitrospirae that inhabit sand, with the addition of a granular media such as anthracite [19]. This method is favored for its speed and reliability in large-scale water treatment facilities. Another filtration method is granular activated carbon filtration (GACF), which is effective in removing organic pollutants. GACF media, typically made from organic materials such as coal and wood, are high in carbon and extremely porous. Contaminants are captured through adsorption on the surface of the granules, providing a habitat for microbial communities, predominantly Polaromonas and Hydrogenophaga species [33]. This dual action of adsorption and microbial degradation makes GACF a versatile and efficient treatment option. Slow sand filtration (SSF) is a more traditional technology where water percolates at a low flow rate through a sand bed layered atop gravel for structural support. A flow regulator is commonly used to maintain consistent flux. This method removes harmful bacteria, pathogens, and turbidity from water. A defining feature of SSF is the formation of a biological layer, known as the schmutzdecke, which develops on the sand’s surface. This layer hosts diverse microbial communities that play a critical role in breaking down contaminants [34]. Each of these biofiltration technologies relies on the interplay of microbial ecology and physical processes. A deeper understanding of the microbial communities and their interactions with the filter media can lead to improved designs and more efficient water treatment solutions.

2.2. Biofiltration of PAHs

Studies have reported specialised methods for separation on PAHs in aqueous sources [35]. Membrane filtration is a method that utilises a barrier material that is able to resist fluid pressure and perform separation of harmful constituents (Figure 5). There are a wide range of methods that filter membranes adopt in the removal of PAHs, such as reverse osmosis [36], ultrafiltration [37], microfiltration, and nanofiltration [5]. It has been shown that waste water can be cleaned with an efficiency of up to 85% through nanofiltration [38] and a separation of up to 72% through reverse osmosis [36]. These methods are all pressure-driven filtration methods and the main difference between them lies in the pore sizes, with microfiltrarion being least selective (largest pores size) and the most selective method of reverse osmosis being able to remove dissolved salts and ions. One study investigates aerated biofiltration as a method to simultaneously remove iron and PAHs, specifically naphthalene, from groundwater. The hypothesis of the work is that manufactured gas plant (MGP) groundwater can be purified by passing the water through a system of columns that have internal gravel media combined with filtered air delivery. This single-stage treatment has significant results for the separation of iron and PAHs, with 99% removal of two- and three-rings PAH compounds. This technology therefore holds promising remediation capabilities [39]. A case study in Compans, France, was performed on a drainage system of a trafficked highway situated in an industrial area that hosts treatment facilities, polymer fabrication, and hydrocarbons and gas storage. Two linear biofilters were studied—a bio-filter swale (BFS) and a vegetative filter strip (VFS)—both positioned in order to receive runoff water from the highway. It was noticed that lower-molecular-weight PAHs were harder to filter (47–88%). However, both systems performed efficiently in removal with high concentration reduction (Ec50) observed (91% for both systems for ${\Sigma }_{16}\mathrm{PAH}$) [40]. Hybrid strategies are adopted for PAH removal, which require ultrafiltration in order to use sand for the removal of coking wastewater and leachate, achieving 53.2% and 54.4%, respectively. PAH reductions from coke on bed particles in cylinder containers incorporating three layers of gravel were recorded to be 36.8% for three-ring compounds, 63.9% for four-ring compounds, 65.2% for five-ring compounds, and 48.4% for six-ring compounds [38]. For municipal leachate these percentages were found to be 52.6% for three-ring compounds, 57.1% for four-ring compounds, 60.6% for five-ring compounds, and 67.7% for six-ring compounds [41]. Moreover, it was reported that layering multiple membranes is a successful strategy in PAH separation, suggesting that a systematic approach to biofiltration has potential for filtration efficacy [42].

2.3. Membrane Filtration of PAHs Through Fungi

While the use of fungal biomass for biosorption and biodegradation of PAHs is well established, the concept of employing fungal mycelial membranes as filtration media specifically for PAH removal in flow-through systems remains under-explored in the peer-reviewed literature (

Table 3. Comparative overview of relevant fungal biofilm and mycelial membrane studies with possible implications for PAH removal applications.

Fungus/Matrix	Media	Contaminant	Effectiveness	Relevance to PAH Filtration	References
Ganoderma sp. dried fungal membrane	Cross-flow filtration	Pb2+	85–90 %	structural analogue	Parasnis et al., 2024; [50]
Mixed mycelium	Mycofiltration bed (e.g., wood chips)	Bacteria, pathogens	>95	functional model	Chandra et al., 2022; [43] Mehta et al., 2017; [44]
Coriolus versicolor fungal membrane	Synthetic waste water	97% TOC and 99% color removal	Textile dye	structural analogue/potential for PAH capture	Hai et al., 2006; [48]
Aspergillus carbonarius fungal membrane	Synthetic waste water	Textile dye	91% decolorisation and 73.2% COD removal	structural analogue/potential for PAH capture	Isik et al., 2019; [49]

). Membrane filtration techniques are usually employed in cases of matter separation, and since PAHs are hydrophobic molecules, they tend to pass through the membrane, unless there is adsorption onto the surface. Moreover, most PAH remediation studies focus on soil/sediment mycoremediation or slurry-phase reactors because they align better with fungal ecology. However, related work in mycofiltration has shown promising results in water treatment. This method typically employs hybrid consortia of fungal mycelium combined with substrates such as wood chips or agricultural grains. This approach has been successfully applied for the removal of harmful waterborne microbes [43,44] and heavy metals [45,46]. In these cases, mycelial mats function as living filtration media capable of retaining microbial pathogens, excess nutrients, and heavy metals from stormwater and wastewater, suggesting their broader potential for contaminant removal in environmental remediation systems [47]. Additionally, the fungal membranes method has been exploited for biofiltration, focusing on treating textile waste water [48,49]. A more targeted example is the employment of a dried mycelium membrane from the genus Ganoderma. The membrane (approximately 5–6 mm thick) was used in cross-flow filtration, which achieved 85–90% removal efficiency of lead (${\mathrm{Pb}}^{2+}$), even at high flux rates. This was achieved by combining mechanical retention and biosorption via functional groups on hyphal surfaces [50]. Although these studies focus on heavy metal contaminants, the architectural features of fungal membranes—porosity, large surface area, and the presence of binding moieties—suggest potential applicability to hydrophobic organic pollutants such as PAHs. Moreover, research on PAH transmembrane transport into fungal cells, particularly in white-rot fungi such as Phanerochaete chrysosporium, indicates that passive diffusion across the cell envelope is a prerequisite for enzymatic degradation, underscoring the fungal membrane’s central role in pollutant uptake [51]. Together, these lines of research lend plausibility to the hypothesis that living fungal membranes could serve as biofilters. Furthermore, in an advanced material context, they could be engineered to provide initial physical retention of PAHs followed by optimised enzymatic transformation.

3. Biosorption and Other Methods

3.1. Overview of Biosorption

Biosorption is a passive physiochemical process which happens in specific biomasses that have functional groups in the cell wall, such as carboxyl, hydroxyl, amino, phosphate, and sulfhydryl [52]. It is a subcategory of adsorption that does not rely on metabolic activity [53]. This method is useful when conventional biodegradation processes are ineffective against particular contaminants; in such cases, the contaminants are instead accumulated and collected on the surface of the biosorbents [54]. On the one hand, some biosorbents are specially engineered for biosorption; these materials often fall into the category of more expensive options due to the precision and resources required for their development. On the other hand, one of the most compelling advantages of biosorption technology lies in its ability to utilize by- or waste products and low-cost materials, making it both economically feasible and widely accessible. For instance, waste materials such as seaweed, eggshells, and cotton have been reported as effective biosorbents [54]. Additionally, industrial by-products like yeast residues from large-scale fermentation processes have demonstrated significant biosorptive potential, offering a sustainable solution for pollutant removal [55]. The ability to source biosorbents from either low-cost waste materials or specialized engineered materials demonstrates biosorption’s adaptability as a treatment technology. It allows for both high-performance applications, where specificity is critical, and large-scale, cost-effective solutions that capitalise on waste valorisation. The attachment to the surface develops through different mechanisms. Ion exchange is found to have a predominant role between the sorbate and the sorbent biomass. One determining factor is the pH and ion concentration defining the binding sites of the ions and influencing the attachment [56]. Another advantage of biosorption is that it is proven to be a reversible process. By adjusting the pH and/or ion concentration, mechanisms such as competition or deprotonation are induced, enabling the regeneration of the biosorbent [57]. Another mechanism of attachment is surface complexation, in which metal ions form chemical bonds or precipitates through reactions at the interface between sorbents and sorbates. This process involves the accumulation of complexes on the sorbent surface. Temperature can have an impact on biosorption: a moderate increase can promote physical adsorption or speed up chemical binding, whereas a substantial rise may impair biosorbent structures, reduce sorption efficiency and raise operational costs [58]. Biosorption serves as the crucial initial step in the broader process of bioremediation, laying the foundation for subsequent bioaccumulation. While these two processes are closely interlinked, they remain distinct in their mechanisms and roles. Biosorption, as previously mentioned, involves the passive binding of pollutants to the surface of the biosorbent, facilitated by functional groups on the cell wall. In contrast, bioaccumulation refers to the active uptake of pollutants into the cell, a process that requires the metabolic engagement of the biomass. Within the cell, these pollutants are transformed or degraded, ultimately contributing to the detoxification and remediation of the environment. Biosorption modulates the efficacy of bioaccumulation; understanding the biosorptive potential of a material is therefore paramount to establishing the feasibility of subsequent bioaccumulation processes. Studies have been conducted on biosorption using fungal pellets, demonstrating success across a broad spectrum of pollutants, including phenolic compounds, pesticides, and heavy metals [52,59,60]. Fungal pellets represent the form in which fungi grow in liquid culture, offering a structure that facilitates biosorption processes. Most research in this area has focused on the use of non-viable or dead fungal cells due to several practical advantages. Firstly, this method eliminates the need for a continuous supply of nutrients during sorption, simplifying operational requirements. Secondly, their deactivation allows the biosorbent to be stored for extended periods without degradation, enhancing its usability and shelf life [52].

3.2. Biosorption of PAHs

The biosorption of PAHs is a complex physicochemical process largely driven by hydrophobic interactions, $\pi –\pi$ stacking, van der Waals forces and the interaction between aromatic PAHs and functional groups present on the surface of biological materials [61]. The pH plays a critical role in modulating the ionization of surface functional groups such as carboxyl and amino groups on biosorbents, which can influence the sorption of charged contaminants like heavy metals [62]. Unlike heavy metals, which exhibit consistent ionic properties, the functional groups of organic pollutants vary according to their chemical structure, resulting in differing degrees of ionization and charge behavior under varying pH conditions [11]. However, for hydrophobic organic compounds such as phenanthrene, biosorption by inactivated bacterial biomass primarily occurs through a partitioning mechanism into hydrophobic domains of the cell wall, rather than via electrostatic interactions or ion exchange [63]. Temperature affects biosorption kinetics and thermodynamics; moderate increases in temperature can enhance the mobility of PAHs and increase sorption capacity, although excessive heat may destabilize biosorbent structures [64]. Biosorbents such as biochar, algae, and microbial biomass exhibit varying affinities for PAHs based on their surface area, porosity, and carbon content, making them promising candidates for environmentally friendly remediation technologies. A variety of sorbents have been tested and reported to be effective. Activated carbon is one of the technologies that is used to remove PAHs and its efficiency is dependent on the type of carbon, contact time, surface area, temperature and PAH concentration [65]. Positive results were reported by a study were six carcinogenic PAHs listed by the Environmental Protection Agency (EPA) were mixed with distilled water and tested against three sorbents: activated carbon WG-12, mineral sorbent and quartz sand. The results present an overall decrease in concentration, with particular efficiency using activated carbon, achieving 100% removal for benzo[ghi]perylene (BghiP), 99.8% for benzo[k]fluoranthene (BkF), 97.1% for benzo[a]pyrene (BaP), 96.7% for benzo[b]fluoranthene (BbF) and dibenzo[ah]anthracene (DahA), and 90.9% for indeno[1,2,3-cd]pyrene (IP). Mineral sorbent performed the least effectively [41]. Even though active carbon is commercially available and performs well in removal, it is not commonly used due to its high cost compared to biomass waste [66]. A study testing soybean stalk-based carbons against removal of phenanthrene (PHE), naphthalene (NAP), and acenaphthene (ACE) in aqueous solution indicates that the increase in carbon quantity and carbonisation temperature is proportional to removal. The best results appear when the preparation of stalk-based carbons is set by phosphoric acid activation at a carbonisation temperature of 700 C with a carbon concentration of 0.04 g per 32 mL. The removal efficiencies observed are 99.89% for PHE, 100% for NAP and 95.64% for ACE. The study identifies opportunities to use biomass waste for sorption purposes [67]. Another economically convenient alternative to activated carbon is activated coke [68]. To assess the adsorption properties of activated coke, petroleum coke was prepared by potassium hydroxide (KOH), before being crushed and submerged in an aqueous solution containing naphthalene (NAP), fluorene (FLU), phenanthrene (PHE), pyrene (PYR) and fluoranthene (FLA). Over 30 min, a removal through sorption of 92.6% for FLA, 93.4% for PYR, 95.2% for PHE, 99.1% for NAP, and 99.7% for FLU occurs. Moreover, the adsorption is correlated to the molecular weight of the contaminants as NAP, suggesting that there is a process of diffusive mass transfer-control, which happened faster with the initial concentration [69]. Overall, activated carbon is an effective method for PAH sorption, but the cost of operations and materials is quite high [70], leading to a search for solutions in waste materials as alternatives.

3.3. Fungal Biosorption of PAHs

Fungal biosorption of PAHs operates through a combination of physico-chemical interactions between fungal biomass and hydrophobic organic pollutants, offering an effective non-metabolic pathway for pollutant removal. The structural composition of fungal cell walls, including chitin, glucans, and melanin, hosts a variety of functional groups such as carboxyl (−COOH), hydroxyl (−OH), amino (−NH₂), sulfonate (−SO₃⁻), and phosphonate/phosphate (PO₄³⁻), which act as active sorption sites [71,72]. These groups mediate PAH binding via electron donor–acceptor mechanisms and planar $\pi –\pi$ interactions with the aromatic rings of PAHs [73]. In natural environments, combined pollution is common, with the co-occurrence of PAHs and heavy metals frequently observed. The simultaneous uptake of these contaminants is often mutually influenced, as the presence of one can alter the biosorption behaviour of the other. In one study, it has been showen that Cu²+ enhances anthracene partitioning into fungal biomass, while Mn²+ inhibits it [74]. In multi-solute scenarios, adsorption sites on fungal biomass are shared among coexisting PAHs and other organic compounds, leading to competitive adsorption. This competition can effect the uptake of single-solute and alters sorption isotherms, for instance when dissolved organic matter or surfactants occupy or block $\pi –\pi$ interaction sites. Co-solutes can, in fact, occupy the same hydrophobic/$\pi$ sites or change apparent PAH solubility. Dissolved organic matter (DOM) and non-ionic surfactants can decrease the sorption of PAHs onto solid phases by competing for adsorption sites and by promoting micellar or colloidal solubilisation that maintains PAHs in the aqueous phase [75]. Moreover, multi-solute PAH systems can also show direct competition for limited aromatic domains (e.g., binary PHE/NAP systems), typically lowering single-solute capacities and changing isotherm nonlinearity [76]. PAH sorpiton to fungal mass can be observed to be near-linear and partition-dominated, Freundlich $N\approx 1$ [77]. As the concentration increases, the functional groups are chemically affected and nonlinearity is predominant with either Freundlich n < 1 or Langmuir models. Pseudo-second-order (PSO) kinetics are typically used to describe the uptake of PAH onto fungal biomass. Anthracene adsorption onto ectomycorrhizal mycelia was well described by PSO kinetics, yielding rate constants of 4.13–6.27 $\mathrm{mg}{\mathrm{g}}^{-1}$ and equilibrium capacities of 1.32–1.79 $\mathrm{mg}{\mathrm{g}}^{-1}$ under set conditions of 25 °C and pH 6 [78]. Phenanthrene adsorption onto P. chrysosporium likewise conformed to PSO behavior, with equilibrium capacities ranging from 0.99 to 2.04 $\mathrm{mg}{\mathrm{g}}^{-1}$ depending on the chemical modification of surface functional groups [77]. PSO constants vary in form and are sensitive to initial concentration ($C_0$) and modeling convention; in general, the rate constant K is not an intrinsic constant and decreases with increasing $C_0$ [79]. Fungal phenolic polymers and melanin are found to have large potential metal-binding sites with oxygen-containing groups [80,81]. Hence, due to good efficiency proprieties and adsorption versatility, fungal biomass has attracted attention as a viable alternative to traditional biosorption technologies [82]. Unlike bacteria sorption that relies on surface sorption, fungal sorption is predominantly governed by partitioning, where PAHs are physically stored within the biomass. Sorption isotherms that help model and predict the sorption capacity and behavior of materials such as naphthalene, acenaphthene, fluorene, phenanthrene and pyrene have been shown to be linear and non-competitive, with Freundlich N values approximating 1, indicating that partitioning, rather than surface saturation, is the dominant mechanism.

The carbon-normalised partition coefficients (Koc) of PAHs to white-rot fungi correlate linearly with the octanol–water partition coefficient (Kow), enabling predictive modeling of biosorption behavior. Quantitatively, partition coefficients (Kd) for different PAHs range from 361.0 mL/g for naphthalene to 24,140 mL/g for pyrene [83]. Additionally, exopolymeric substances (EPS) surrounding fungal hyphae enhance biosorption by physically trapping PAHs and localising lignin-degrading enzymes such as laccase and manganese peroxidase, which can further degrade adsorbed compounds [84]. Inactivated fungi often exhibit greater sorption capabilities compared to active ones, for example, In the case of BaP adsorption efficiency of inactive fungi (40.76%) is greater than live fungi (34.24%) [72]. Nevertheless, live fungi combine biosorption with biodegradation, storing up to 60–85% of PYR and 40–65% of PHE in fungal biomass prior to enzymatic breakdown [83]. Biosorption of fungi is a multifaceted process in PAH remediation, with fungi functioning as both passive sorbents and active bioreactors under diverse environmental conditions.

4. Biodegradation and Other Methods

4.1. Overview of Biodegradation

Biodegradation physical treatment intervention is most commonly categorized into the following categories: in situ and ex situ [85]. Ex situ techniques consist of excavation and transport of contaminated masses to specialized treatment or disposal facilities. In-situ methods seek to isolate or remediate contaminants in the field [86] (Figure 6). A wide range of methods are employed for the removal of contaminants, depending on the nature and extent of pollution. These include physical, thermal–chemical and biological treatments. Physical treatments involve a physical separation to remove contaminants, such as soil washing, capping, air stripping, sedimentation, and filtration [7]. There are also other extraction techniques involving solvent extraction, vacuum extraction, and electrokinetic remediation, which draw contaminants out of the affected material [87]. Thermo-chemical techniques utilize thermal processes and chemical reactions to neutralise contaminants, including thermal desorption [88], incineration [89], pyrolysis [90], chemical oxidation [91], precipitation and neutralization [92]. Biological treatment harnesses microbial activity for pollutant degradation utilizing for example bioreactors and biopiles [85]. Meanwhile, application of ex situ pollutant remediation poses a high risk while transporting and handling the pollution; it has a high cost and there is a high risk of exposure to contaminants. In situ remediation processes require a longer period of time for completion but make use of more accessible and low-cost technologies. However, they may be less efficient than the ex situ methods and need careful site assessment, monitoring and risk management [86]. One of the most promising strategies in the in-situ category is bioremediation, as it uses biological factors in order to remediate the environment. Leveraging the metabolic process of living organisms is an alternative approach to environmental clean-up by utilising their natural ability to degrade, transform, or remove pollutants from contaminated sites. It offers a cost-effective and environmentally friendly alternative to traditional remediation methods [7]. The strategy leverages the metabolic pathways of organisms to break down and remove hazardous substances, including hydrocarbons, heavy metals and pesticides (Table 4 and Table 5). In microbial biodegradation, certain bacteria, fungi, or even algae are selected for their ability to degrade specific toxins (PAHs) (Figure 7). Different microbial communities have distinct optimal pH ranges: many bacteria prefer neutral to slightly alkaline conditions, while fungi can tolerate more acidic environments. A similar principle applies to temperature, as optimal temperatures can enhance the degradation of specific pollutants [93]. The most commonly employed in situ bioremediation strategies are biostimulation and bioaugmentation. When nutrient availability is limited, biostimulation becomes a key approach. It targets areas that are deprived and aims to stimulate indigenous organisms by adding supplementation of nutrition through phosphorous, nitrogen and carbon sources, providing favourable environmental conditions to promote growth. Furthermore, if the indigenous microflora lack biocompatibility to degrade pollutants, it is possible, through bioaugmentation, to introduce specialised organisms that are provided with a compatible metabolic pathway that enables degradation. As these microorganisms proliferate, they break down the pollutants into less harmful compounds through processes such as biodegradation, biotransformation, or mineralization. Biostimulation and bioaugmentation techniques can be used separately or in combination, depending on the site, the environmental conditions and the pollutants that are targeted [94]. Although these methods are human-driven interventions, they are generally considered safe, as they enhance nutrient availability in ways that mimic natural processes. The treatments themselves do not persist or accumulate in the environment. Moreover, in cases where external microorganisms are introduced, they are typically selected from species closely related to the native microbiota and their populations tend to decline naturally once the pollutant source is depleted [95]. Extended studies have been conducted on fungal and bacterial species with regard to their natural ability for hydrocarbon detoxification [96,97,98,99,100,101].

Figure 6. Bioremediation strategies ex and in situ.

Figure 6. Bioremediation strategies ex and in situ.

Figure 7. Pathways for PAH degradation by bacteria and fungi. Based on [102]; figure created by the authors.

Figure 7. Pathways for PAH degradation by bacteria and fungi. Based on [102]; figure created by the authors.

Table 4. Biological agents and processes for pollutants bioremediation: effectiveness and references.

Pollutant	Organism	Enzyme/Process	Effectiveness	References
Benzo[a]pyrene, Pyrene, Chrysene, Fluoranthene, Naphthalene, Phenanthrene, Anthracene	Mycobacterium gilvum MI, Novosphingobium pentaromativorans sp. nov	Ring-hydroxylating dioxygenases (RHDs), Dehydrogenases, Monooxygenases	88.2–99.9 % after 8 days	Sohn et al., 2004 [103]
Phenanthrene, Pyrene, Benzo[a]pyrene, Benzo[b]fluoranthene	Mycobacterium gilvum MI, Mycobacterium sp. ZL7, Rhodococcus rhodochrous Q3	Ring-hydroxylating dioxygenases (RHDs), Dehydrogenases, Monooxygenases	consortium H6 (Q3:ZL7:MI = 1:2:2) 59% within 8 days	Zhou et al., 2023 [104]
Pyrene	Pseudomonas aeruginosa strain ASU-B6	not specified	92% after 15 days	Mawad et al., 2024 [105]
Hexavalent chromium (Cr(VI))	Geobacter sulfurreducens	Extracellular protein-mediated reduction; intracellular accumulation	99% of 100 mg/L Cr(VI) using a cell density of 5.8 × ${10}^8$ cells/mL. 99% of 200 mg/L Cr(VI) using 11.4 × ${10}^8$ cells/mL.	Elmeihy et al., 2021 [106]
Cadmium (${\mathrm{Cd}}^{2+}$)	Pseudomonas aeruginosa	Bioaccumulation and Biosorption	Up to 94.7%	Chellaiah, 2018 [107]
Hexavalent chromium (Cr(VI))	Priestia megaterium strain BM.1	Bioreduction and adsorption on hydrochar	97% removal of Cr(VI) (initially 60 mg/L)	Wu et al., 2025 [108]
Pentachlorophenol (PCP)	Sphingomonas chlorophenolica	Pentachlorophenol hydroxylase, Tetrachlorohydroquinone dehalogenase, 2,6-Dichlorohydroquinone dioxygenase	Not specified	Copley, 2000 [109]
Perchloroethylene, Trichloroethylene, Dichloroethene, Vinyl chloride	Dehalococcoides spp	Reductive dehalogenases (RDases)	Not specified	Vainberg et al., 2009 [110]
Anthracene, Acenaphthene, Fluoranthene, Fluorene	Aspergillus niger	Reductive dehalogenases (RDases)	ter 30 days: 77.8%, 65%, 60.9%, 52.5%	Manjunatha et al., 2025 [111]
Phenanthrene, Pyrene	Podoscypha elegans FTG4	Laccase, Lignin peroxidase (LiP), Manganese peroxidase (MnP)	50.6% of PHE, 48% of PYR, for 50 mg/kg, 99% of PHE, 98.9% of PYR for 20 mg/L	Agrawal et al., 2021 [112]
Zinc ion (${\mathrm{Zn}}^{2+}$), Chromium ion (CR3+), Lead ion (Pb2+)	Phanerochaete chrysosporium, Trametes versicolor	biosorption	P. c adsorbs Cr3+ medium at concentrations of 0.5, 1 mg ${\mathrm{L}}^{-1}$. T. v adsorbs Pb2+ at concentrations of 0.25, 1, 2 mg ${\mathrm{L}}^{-1}$	Solis Pacheco et al., 2015 [113]

Table 4. Biological agents and processes for pollutants bioremediation: effectiveness and references.

Pollutant	Organism	Enzyme/Process	Effectiveness	References
Benzo[a]pyrene, Pyrene, Chrysene, Fluoranthene, Naphthalene, Phenanthrene, Anthracene	Mycobacterium gilvum MI, Novosphingobium pentaromativorans sp. nov	Ring-hydroxylating dioxygenases (RHDs), Dehydrogenases, Monooxygenases	88.2–99.9 % after 8 days	Sohn et al., 2004 [103]
Phenanthrene, Pyrene, Benzo[a]pyrene, Benzo[b]fluoranthene	Mycobacterium gilvum MI, Mycobacterium sp. ZL7, Rhodococcus rhodochrous Q3	Ring-hydroxylating dioxygenases (RHDs), Dehydrogenases, Monooxygenases	consortium H6 (Q3:ZL7:MI = 1:2:2) 59% within 8 days	Zhou et al., 2023 [104]
Pyrene	Pseudomonas aeruginosa strain ASU-B6	not specified	92% after 15 days	Mawad et al., 2024 [105]
Hexavalent chromium (Cr(VI))	Geobacter sulfurreducens	Extracellular protein-mediated reduction; intracellular accumulation	99% of 100 mg/L Cr(VI) using a cell density of 5.8 × ${10}^8$ cells/mL. 99% of 200 mg/L Cr(VI) using 11.4 × ${10}^8$ cells/mL.	Elmeihy et al., 2021 [106]
Cadmium (${\mathrm{Cd}}^{2+}$)	Pseudomonas aeruginosa	Bioaccumulation and Biosorption	Up to 94.7%	Chellaiah, 2018 [107]
Hexavalent chromium (Cr(VI))	Priestia megaterium strain BM.1	Bioreduction and adsorption on hydrochar	97% removal of Cr(VI) (initially 60 mg/L)	Wu et al., 2025 [108]
Pentachlorophenol (PCP)	Sphingomonas chlorophenolica	Pentachlorophenol hydroxylase, Tetrachlorohydroquinone dehalogenase, 2,6-Dichlorohydroquinone dioxygenase	Not specified	Copley, 2000 [109]
Perchloroethylene, Trichloroethylene, Dichloroethene, Vinyl chloride	Dehalococcoides spp	Reductive dehalogenases (RDases)	Not specified	Vainberg et al., 2009 [110]
Anthracene, Acenaphthene, Fluoranthene, Fluorene	Aspergillus niger	Reductive dehalogenases (RDases)	ter 30 days: 77.8%, 65%, 60.9%, 52.5%	Manjunatha et al., 2025 [111]
Phenanthrene, Pyrene	Podoscypha elegans FTG4	Laccase, Lignin peroxidase (LiP), Manganese peroxidase (MnP)	50.6% of PHE, 48% of PYR, for 50 mg/kg, 99% of PHE, 98.9% of PYR for 20 mg/L	Agrawal et al., 2021 [112]
Zinc ion (${\mathrm{Zn}}^{2+}$), Chromium ion (CR3+), Lead ion (Pb2+)	Phanerochaete chrysosporium, Trametes versicolor	biosorption	P. c adsorbs Cr3+ medium at concentrations of 0.5, 1 mg ${\mathrm{L}}^{-1}$. T. v adsorbs Pb2+ at concentrations of 0.25, 1, 2 mg ${\mathrm{L}}^{-1}$	Solis Pacheco et al., 2015 [113]

Table 5. Biological agents and processes for pollutant bioremediation: effectiveness and references.

Pollutant	Organism	Enzyme/Process	Effectiveness	References
Cadmium (II), manganese (II), zinc (II) ions	Ganoderma lucidum heteropolysaccharides (GLHP)	biosorption	98.2%, 80.6%, 82.8%	Marolt et al., 2024 [114]
Cadmium	Aspergillus niger	biosorption	82.2%	Amini et al., 2009 [115]
Carbamates	Ascochyta sp. CBS 237.37	not specified	36–94.8%	Kaur and Balomajumder, 2019 [116]
Parathion, Terbufos, Azinphos-methyl, Phosmet, Tribufos, Trichlorfon	Bjerkandera adusta 8258, Pleurotus ostreatus 7989, Phanerochaete chrysosporium 3641	Cytochrome P450 monooxygenases	50–96%	Jauregui et al., 2003 [117]
Methyl parathion (O,O-dimethyl-O-4-nitrophenyl phosphorothioate)	Chlorella vulgaris, Scenedesmus obliquus, Nannochloropsis spp., Anabaena spp., Spirulina spp.	electrostatic attraction and complexation	24–98%	Satpati et al., 2023 [118]
Naphthalene, Phenanthrene, Anthracene, Fluoranthene, Benzo[a]pyrene	Aspergillus niger	biosorption	82.2%	Amini et al., 2009 [115]
Phenol	Chlorella sp., Scenedesmus sp., Spirulina sp., Chlamydomonas sp.	Phenol hydroxylase, Catechol 1,2-dioxygenase (C12O), Catechol 2,3-dioxygenase (C23O), Laccases and peroxidases, Polyphenol oxidase	Chlorella sp.: up to 90%, Scenedesmus sp.: 70–85%	Al-Dahhan et al., 2018 [119] Radziff et al., 2021 [120]

Table 5. Biological agents and processes for pollutant bioremediation: effectiveness and references.

Pollutant	Organism	Enzyme/Process	Effectiveness	References
Cadmium (II), manganese (II), zinc (II) ions	Ganoderma lucidum heteropolysaccharides (GLHP)	biosorption	98.2%, 80.6%, 82.8%	Marolt et al., 2024 [114]
Cadmium	Aspergillus niger	biosorption	82.2%	Amini et al., 2009 [115]
Carbamates	Ascochyta sp. CBS 237.37	not specified	36–94.8%	Kaur and Balomajumder, 2019 [116]
Parathion, Terbufos, Azinphos-methyl, Phosmet, Tribufos, Trichlorfon	Bjerkandera adusta 8258, Pleurotus ostreatus 7989, Phanerochaete chrysosporium 3641	Cytochrome P450 monooxygenases	50–96%	Jauregui et al., 2003 [117]
Methyl parathion (O,O-dimethyl-O-4-nitrophenyl phosphorothioate)	Chlorella vulgaris, Scenedesmus obliquus, Nannochloropsis spp., Anabaena spp., Spirulina spp.	electrostatic attraction and complexation	24–98%	Satpati et al., 2023 [118]
Naphthalene, Phenanthrene, Anthracene, Fluoranthene, Benzo[a]pyrene	Aspergillus niger	biosorption	82.2%	Amini et al., 2009 [115]
Phenol	Chlorella sp., Scenedesmus sp., Spirulina sp., Chlamydomonas sp.	Phenol hydroxylase, Catechol 1,2-dioxygenase (C12O), Catechol 2,3-dioxygenase (C23O), Laccases and peroxidases, Polyphenol oxidase	Chlorella sp.: up to 90%, Scenedesmus sp.: 70–85%	Al-Dahhan et al., 2018 [119] Radziff et al., 2021 [120]

4.2. Biodegradation Against PAHs Through Bacteria

Microbial activity plays a primary role in the degradation process; the compounds are transformed into metabolites that are less complex and harmful for the environment [12]. Various microorganisms can degrade PAHs, including specialized prokaryotic bacteria and eukaryotic organisms such as fungi. Bacteria have proven to be very effective in the remediation of petrochemical waste. The primary mechanism of PAH degradation by bacteria relies on enzyme-driven activation and oxidation, facilitating oxygen fixation [121]. BaP is considered the most dangerous and carcinogenic and Sphingomonas paucimobilis is found to be effective in its degradation. In one study, a decrease of BaP of 5% in liquid culture was shown after 168 h [122]. From a polluted site, Aitken et al. were able to collect 11 strains of indigenous bacteria with the ability to degrade BAP, including 3 species belonging to the genus Pseudomonas, as well as Agrobacterium, Bacillus, Burkholderia and Sphingomonas species [123]. It has been shown that soil with a high quantity of PAHs contains more bacteria per gram compared to soil with a lower amount of PAHs. Using already specialised bacteria that are indigenous is a favourable approach to decontamination and several of these organisms use contaminants as sources of energy and carbon [124]. Pseudomonas sp., Alcaligenes sp., Rhodococcus sp., Beijerinckia sp. are reported to be able to degrade high molecular weight PAHs that have three or more benzene rings [125]. A limiting factor in the context of bacteria remediation is accessibility to oxygen for bacteria that need aerobic condition to survive; for this reason, techniques such as bioventing with injection of oxygen are more suitable in this context, or alternatively allowing waste water to enter the plant while being aerated [126].

4.3. Fungal Biodegradation Against PAHs

Some fungi possess the ability to thrive in harsh environments, whether terrestrial or aquatic, making them ideal candidates for bioremediation. Psathyrella species for example are able to live in arctic environments and in the presence of toxic pollution [127]. The use of fungi for remediation is called mycoremediation [128]. White rot fungi are a distinct group of fungi that are recognized to degrade a wide range of pollutants, including PAHs, through their enzyme production for lignin degradation [129]. What makes this technology particularly appealing is its cost-effectiveness (

Table 6. Comparison of bioremediation treatment costs [130].

Treatment	Approximate Cost (£25/Tonne Soil)
Biological	5–170
Chemical	12–600
Physical	20–170
Solidification/stabilisation	17–171
Thermal	30–750

) [130]. Unlike bacteria, which often require complex and costly nutritional sources, fungi can flourish on substrates sourced from agricultural waste, making their application both practical and sustainable [131]. Fungi have been successfully utilised in remediating both soil and water contamination. Fungal diversity in wastewater treatment systems enhances efficiency across a range of industries such as pharmaceuticals and it is present in textile and distillery activities, demonstrating the strong potential of fungi as a powerful tool for tackling industrial pollution [132]. Fungal PAH metabolism can be classified into two major types: those mediated by non-ligninolytic fungi and those mediated by ligninolytic fungi [133]. The initial step in PAH metabolism by non-ligninolytic fungi involves the oxidation of the PAH through a cytochrome P450 monooxygenase enzyme-mediated reaction, resulting in the formation of an arene oxide [134]. White rot fungi harness the power of their lignin-modifying enzymes (LMEs) to break down lignin during wood decomposition. The key enzymatic groups; lignin peroxidases (LiPs), manganese-dependent peroxidases (MnPs), and laccases, are not only pivotal in natural wood degradation but also demonstrate extraordinary versatility in environmental remediation [135]. These enzymes enable white rot fungi to degrade a broad spectrum of organic pollutants, including persistent contaminants like polychlorinated biphenyls (PCBs), polycyclic aromatic hydrocarbons (PAHs), pesticides, synthetic dyes, and various chlorinated and phenolic compounds. Most recently, studies have reported enzymes such as feruloyl esterase (EC 3.1.1.73), aryl-alcohol oxidase (EC 1.1.3.7), quinone reductases (EC 1.6.5.5), lipases (EC 3.1.1.3), xylanase (EC 3.2.1.8) and catechol 2, 3-dioxygenase (EC 1.13.11.2) that indirectly facilitate the degradation of pollutants [136,137]. LiPs directly oxidize specific PAHs, making ligninolytic fungi unique among eukaryotes in their ability to break down fused-ring aromatics, a process once believed to be exclusive to bacteria [138]. This enzyme is able to target and break C-C and C-O bonds in lignin but also plays a crucial role as a catalyst for the oxidation of several non-aromatic and aromatic compounds; therefore, it allows the degradation of pollutants such as PAHs and many more, like phenolic compounds and textile dyes. Manganese Peroxides (MnP), like LiP, are heme-containing enzymes; they possess a high level of oxidation towards Mn(II) to Mn(III). Mn(III) produces a reactive radical able to tackle the aromatic rings with high molecular weight in PAHs. Laccase is a copper enzyme able to oxidase less complex PAHs molecules with low molecular weight. In order to augment the degradation efficiency of more complex PAHs with more than four rings, a highly hydrophobic and thermally stable electron shuttle is needed to act as mediator in the degradation process [132,139,140]. The overall process of PAH oxidation generates distinct intermediates. It begins with hydroxylation and oxidation of the aromatic rings catalysed by extracellular lignin enzymes or intracellular monooxygenases with production of arene oxides, dihydrodiols and hydroxylated PAHs [141]. These catalyse oxidation of aromatic rings into quinones, dihydrodiols, or diphenols, which are further cleaved into intermediates like catechol or protocatechuate. This process enables fungi such as Phanerochaete chrysosporium and Pleurotus ostreatus to transform and partially mineralize PAHs [102]. The production and mechanisms of enzyme synthesis are highly species-dependent; however, their synergistic interactions enable the remediation of a diverse array of environmental toxins. Pleurotus ostreatus, in low and high nitrogen conditions, is able to mineralise and metabolise PYR, FLU, PHE and ANT [142]. In one study, G. lucidum was found to be effective in PAH remediation. It has been shown to degrade 99.65% of 20 mg/L PHE and 99.58% of PYR in mineral salt broth after 30 days of incubation at 27 °C. In the PHE-containing broth, G. lucidum produced a maximum of 10,788.00 U/L laccase, 3283.00 U/L lignin peroxidase, and 47,444.00 U/L manganese peroxidase. In the PYR-containing broth, the fungus reached maximum enzyme production levels of 10,166.00 U/L laccase, 3613.00 U/L lignin peroxidase, and 50,977.00 U/L manganese peroxidase [96]. A report showed mineralisation after 21 days of 50% of PYR, 68% of ANT and 63% of PHE by Pleurotus ostreatus, with increased degradation of 75%, 80% and 75%, respectively, by adding supplementation of Tween 40 [99]. Schizophyllum commune ability of PAHs degradation is tested in artificial marine water (AMW) solution with concentration of 50 ppm of PHE, PYR and BaP in aerobic and anaerobic treatments. Positive results of 25%, 18% and 13% were found for PHE, PYR, and BaP, respectively, after 10 days under anaerobic conditions [143]. Fungi’s adaptability to various environmental conditions and ability to degrade a wide range of PAHs, even in low-nutrient settings, make them particularly advantageous for bioremediation. Additionally, fungi can penetrate and colonise contaminated substrates, enhancing their efficiency in pollutant degradation. These characteristics position fungi as a valuable tool for the bioremediation of PAH-contaminated environments.

5. Discussion

This review of biofiltration, biosorption and biodegradation allows us to outline a set of design considerations for fungal-based PAH remediation systems. Influent water quality strongly shapes system performance; laboratory studies generally report effective degradation at influent concentrations between 20 and 50 mg/L, particularly for low- to medium-molecular-weight compounds such as phenanthrene and pyrene. In contrast, high-molecular-weight PAHs exhibit lower removal efficiencies under similar conditions, reflecting their limited solubility and bioavailability. Performance can also be constrained in multi-solute scenarios, where adsorption sites are competitively occupied or in waters containing high levels of dissolved organic matter (DOM). Optimal conditions for sustaining fungal metabolism and enzymatic activity include maintaining pH values between 6 and 8 and temperatures in the range of 20–30 °C. With respect to PAH filtration, fungal living membranes remain under-explored. Current applications are either biologically inactive or focus on pollutants other than PAHs. Current approaches involve lignocellulosic carrier matrices that act as both structural supports and nutrient reservoirs, particularly when derived from low-cost agricultural residues. To ensure long-term operation, effluent PAH concentrations must be continuously monitored, while hydraulic performance should be assessed through flux measurements, where significant declines indicate the need for cleaning or replacement. Microbial partnerships play an important role. Co-cultivation of fungi with bacteria has repeatedly been shown to enhance degradation efficiencies, in some cases by 10–50%, by combining initial bacterial transformation with fungal oxidative breakdown. Biological remediation for contaminant capture and degradation, particularly for PAHs, shows considerable promise as a sustainable environmental regenerative solution. The integration of specific aspects from filtration, sorption and degradation through fungal membrane systems presents an opportunity for a comprehensive remediation method (

Table 7. Advantageous aspects to employ fungal membrane in an integrated process of biofiltration, biosorption and bioremediation.

Aspect	Fungal Membrane Filtration	Fungal Biosorption	Fungal Biodegradation
Pollutant range	Potential to filter microbes and organic particles	Binds metals, dyes, hydrocarbons, complex organics	Degrades complex pollutants
Mechanism type	Functional groups on fungal cell walls	Physical entrapment of through the membrane	Extracellular enzymes
Scalability & maintenance	Potential for in locus degradation therefore reduce fouling, improving durability	Depends on production process of membrane. Potential to reuse biomass, low cost	Scalable for liquid and solid phases
Environmental tolerance	Potentially more resilient	Can tolerate harsh conditions	Can tolerate harsh conditions
Substrate growth and cost	Potentially can grow on low-cost substrates and self-regenerating membranes	Low-cost biomass from industrial waste	Can grow on agro-industrial waste
Synergy potential	Potentially can be integrated in a consortia with bacteria	Compatible with integrated remediation	High synergy with bacteria and plants

). While current bioremediation technologies offer selective specialisation, holistic approaches that address the complete lifecycle of contaminants in aquatic environments remain underdeveloped. White-rot fungi emerge as ideal candidates for membrane-based filtration materials due to their naturally occurring entangled mycelial network structures, while their inherent role as decomposers in natural ecosystems facilitates localised degradation capabilities. These characteristics establish a robust foundation for developing systematic remediation processes utilizing fungal mycelium as a multifunctional filtration medium that simultaneously captures, retains, and degrades PAH contaminants. Combining three methods through the production of pure fungal membranes leverages the properties of the biological matrix and constructs the filtration medium directly from selected living organisms, such as fungal mycelium. The process allows for a more precise understanding and control of the microbial ecology within the system. This approach fundamentally transforms common biofiltration methods by eliminating external microbial community deposition on inert filter substrates. Instead, the living fungal matrix itself serves as both the structural filtration medium and the active biological agent. This proactive approach consists of integrating the membrane composition and microbial functionality as fundamental parts of the design of the filtration system. By unravelling the relation between organism and environment conditions, such as contaminants, temperature, pH and identifying the optimal compositions of microbial consortia, the bioremediation design allows for a robust and efficient system. The control over the microbial ecology allows us to deliberately engineer the biological components for targeted pollutant removal. This biocentric methodology offers potential for adaptive responses to changing environmental conditions, as the living matrix can be modified to meet new challenges.

Taken together, these considerations move fungal-based remediation beyond descriptive case studies toward actionable design principles, bridging laboratory research and engineered living water treatment systems. They also provide a foundation for advancing the spectrum of Engineered Living Materials (ELMs). Insights into how biological systems can be harnessed for remediation inform the engineering of ELMs by optimising embedded biological components for performance, adaptability, and self-organization. In this way, PAH remediation serves as both a testbed and a conceptual bridge to single organisms or consortia of multifunctional living materials capable of dynamically interacting with their environments.