Bioremediation utilizes the natural role of microorganisms and/or their enzymes in contaminated media to transform organic pollutants to levels at which they no longer pose risks to humans, animals, or plants.

Studies have shown that more than 200 species of bacteria can biodegrade PHCs, and various bacteria including Pseudomonas, Nocardia, Vibrio, Corynebacterium, Candida, Arthrobacter, Rhodotorula, Brevibacterium, Flavobacterium, Sporobolomyces, Achromobacter, Bacillus, Aeromonas, Thiobacillus, Acinetobacter, Lactobacter, Staphylococcus, Penicillium, Articulosporium, Halomonas, Klebsiella, Proteus, Aspergillus, Micrococcus, Neurospora, Rhizopus, Mucor, and Trichoderma are known to play vital roles in PHC degradation (Xu et al., 2018; Victor et al., 2020). The most commonly used bacterial remediation techniques are bioaugmentation and biostimulation or a combination of these two processes. Table 1 shows the bioremediation of PHC-polluted soil by the different bacterial strains.

Mycoremediation refers to the use of different species of fungi for the remediation of PHC-contaminated soils (Akhtar and Mannanul, 2020). Mycoremediation is a fungal species-assisted process for the complete mineralization of pollutants into CO₂, H₂O, N₂, and microbial biomass after the completion of their metabolic activities (Rhodes, 2014). The unique morphology and diverse metabolic competence of fungi make them imperative biological agents for bioremediation of contaminants (Deshmukh et al., 2016). On the other hand, fungi have already been recognized for effective contaminant elimination and the progress of environmental biotechnology in the 1980s. Fungi are diverse, contain millions of species that are ubiquitous in various environments, and have been applied in diverse areas (Hawksworth and Lücking, 2017). Fungi can survive and grow well in terrific environmental situations through the distribution of spores and the secretion of various enzymes. Hence, considerable research has been conducted on the applied and fundamental aspects of mycoremediation in both soil and water (Anderson and Juday, 2016). Consequently, mycoremediation has become a firm technology for soil remediation without the threat of the spread of pollutants or their bioaccumulation in the food chain. The mycoremediation potential of PHC-polluted soils by the most studied and effective fungal strains is shown in Table 2.

Dhailappan et al. (2022) studied the degradation rate of PHC-contaminated soil using an edible oyster mushroom (Pleurotus ostreatus). Researchers have used paddy husks to stimulate remediation. The maximum TPH degradation rate was 87% after 21 days (Dhailappan et al., 2022). In another study, the degradation efficiency of a fungal consortium isolated from polluted soil was investigated to detoxify petroleum hydrocarbon-polluted soils. The results showed that the fungal community consisting of two isolates of Aspergillus sp. was the best consortium and had the highest potential to degrade hydrocarbons, mainly n-alkane-contaminated soil, at 50.61% in 16 days (Novianty et al., 2021). Naturally occurring microorganisms in soil matrices play a vital role in the mineralization of crude oil-polluted soil. This study investigated the degradation of conventional diesel, heating diesel fuel, synthetic diesel (syntroleum), fish biodiesel, and a 20% biodiesel–diesel blend by fungal communities under favorable environmental conditions. Fungal communities (Trichoderma sp., Penicillium sp. Giberella sp., Mortierella sp., and Fusarium species) remediation was observed with petroleum, which showed the highest total hydrocarbon decontamination rate of 48% in 28 days (Horel and Schiewer, 2020). Daâssi and Almaghribi (2021) explored the remediation of petroleum-hydrocarbon-contaminated soils using fungal species. After 60 days of incubation, the highest rate of TPH degradation was recorded at approximately 92% ± 2.35% by the Paecilomyces formosus strain (Daâssi and Almaghribi, 2021).

Marín et al. (2018) investigated the potential of endophytic fungi isolated from the stem and leaf of a plant in the Ecuadorian Amazon region for the degradation of oil-contaminated soil. A total of 156 endophytic fungi were isolated, and endophyte Xylaria sp. 1 showed the best TPH removal rate of 98.78% by infrared spectroscopy (IR) and gas chromatography 99.8% (GC) in an in vitro test for 30 days (Marín et al., 2018). Essabri et al. (2019) studied the degradation of TPH in petroleum-contaminated soil with fungi isolated from olive oil effluent using bioaugmentation and biostimulation methods. The results showed that the A. niger fungal isolate greatly degraded total petroleum hydrocarbons, with a degradation of approximately 71.19% at the end of the 60 days treatment (Essabri et al., 2019).

The petroleum hydrocarbon mineralization ability of fungi and plants was investigated by Ani et al. (2021), and the results confirmed that different rates of hydrocarbon degradation by isolated fungi were observed. An in vitro biodegradation study showed that Aspergillus niger, Fusarium solani, Curvularia lunata, and Trichoderma harzianum were the most effective in degrading kerosene (78%), diesel (70%), spent engine oil (83%), and crude oil (77%), respectively (Ani et al., 2021).

Kristanti et al. (2011) investigated microbial decomposition of crude oil-contaminated soil using the white rot fungal species Polyporus sp. S133. The strain was extracted from the PH-contaminated soil site and maintained using the inoculum enrichment technique at 4°C before use. The isolated fungal strain was then applied to a sample containing 200 g of soil and 3,000 mg of crude oil. The maximum degradation rate (93%) was obtained after 60 days of inoculation (Kristanti et al., 2011). Similarly, remediation of n-hexadecane hydrocarbons was investigated using the fungal strain Aspergillus sp. RFC-1 was obtained from oil-contaminated soil samples from the Rumaila oilfield in Basra, Iraq. The experimental results revealed that Aspergillus sp. RFC-1 degraded 86.3% of the initial n-hexadecane concentration after 10 days of incubation when used as the carbon source (Al-Hawash et al., 2018b). Another investigation was carried out by (Mohammadi-Sichani et al. (2019) to assess whether PHC-polluted soil was cleaned using spent mushroom compost as an inoculum for the bioremediation of oil-processing plant spills. The observed results showed that all composts of white-rot fungi, A. bisporus, P. ostratus, and G. lucidum, were able to degrade 71.5%, 69.5%, and 57.5% of the PHCs over 3 months. Among these strains, only A. bisporus showed a high ability to degrade TPHs (71.5%) (Mohammadi-Sichani et al., 2019). Its ability to degrade high- and low-molecular-weight PHCs makes spent mushroom compost potentially useful for mycoremediation. In another study, the biodegradation of phenanthrene was explored using the endophytic fungus Phomopsis liquidambari obtained from the interior of the stem bark of Bischofia polycarpa. The strain reduced more than 77% of the initial concentration of 50 mg/L phenanthrene substrate within 10 days of incubation (Fu et al., 2018). In contrast, a synergistic hydrocarbon (HC) remediation experiment using novel rhizospheric fungal strains in combination with potential plant roots was performed at different concentrations of crude oil-contaminated soil (Asemoloye et al., 2017). The strain was isolated from an aged oil-contaminated site and deployed along with a guinea grass root (Megathyrsus Maximus) for 90 days. The strain achieved 95.36% of total PAH mineralization at 30% and 40% contaminated soil sample concentrations. In addition, the synergistic action of fungal consortia and Megathyrsus maximus roots accelerates the hydrocarbon degradation rate and soil remediation and enriches soil nutrients. Evidence has also shown that fungal strains of marine origin, such as Tinctoporellus sp. CBMAI 1061 and Marasmiellus sp. CBMAI 1062, and Peniophora sp. CBMAI 106) effectively detoxifies large-molecular-weight pyrene and benzo [a]pyrene (BaP) (Vieira et al., 2018). Marasmiellus sp. were subjected to pyrene degradation and were able to completely degrade pyrene after 48 h of incubation. Correspondingly, the ligninolytic fungus Ganoderma lucidum CCG1, extracted from the hardwood stump, degraded phenanthrene and pyrene. G. lucidum could degrade 99.6% of phenanthrene and pyrene in a salt medium after 30 days at 27°C. G. lucidum is a strong PAH degrader in contaminated soil environments (Agrawal et al., 2018). Similarly, the white-rot fungus Coriolopsis byrsina strain APC5 decontaminates pyrene under various soil conditions. The highest pyrene degradation (96.1%) was obtained using ligninolytic enzymes at pH 6°C and 25°C. Based on this examination, C. byrsina APC5 strain can also be used as a potential degrader for PAH contaminants from the contaminated soil environment (Agrawal and Shahi, 2017)

This review briefly elucidates the aerobic and anaerobic degradation pathways and mechanisms of n-alkane degradation in soil using bacterial strains. The PHC-saturated fraction is composed of n-alkanes ranging from C₁ to > C₄₀ (Abbasian et al., 2015).

The fungal remediation mechanism in this review elucidates the microbial pathways involved in decontaminating PAH-polluted soil. The mechanisms of detoxification of PAH-polluted soil by fungi are quite different from those of bacterial remediation catabolism pathways. Fungi secrete a series of intracellular and extracellular enzymes such as lignin peroxidase and manganese peroxidase to remove PAHs from the soil (Reyes-César et al., 2014). Based on the group of fungi and secreted enzymes, the degradation of PAH-polluted soil by fungi can be classified as ligninolytic fungi (a group that secretes enzymes to break down lignin in wood) and non-ligninolytic fungi (a group that does not secrete these enzymes) (Marco-Urrea et al., 2015).

White-rot fungi (WRF) and other macrofungi are a major group of fungi that secrete ligninolytic enzymes to degrade lignin in wood and other organic material considered to be ligninolytic fungi (Rathankumar et al., 2022). Ligninolytic enzymes, such as laccases and peroxidases, are produced extracellularly and oxidize PAHs via nonspecific radical-based reactions. The two major types of peroxidase ligninolytic enzymes are manganese peroxidase and lignin peroxidase, which have the potential to oxidize PAHs into non-toxic forms using direct or indirect oxidation (Al-Hawash, 2018). Under ligninolytic conditions, WRF can oxidize PAHs by generating free radicals (hydroxyl free radicals) via donation of one electron, which oxidizes the ring of PAH and converts it into quinones. Overall, ligninolytic fungi mineralize PAHs into quinones and further into inorganic compounds, such as carbon dioxide and water, as shown in Figure 7. On the other hand, the most common PAH-degrading non-ligninolytic fungi are ascomycetes, zygomycetes, hyphomycetes, Aspergillus sp., Cunninghamella, and Penicillium (Rathankumar et al., 2022). The initial step of PAH oxidation by non-ligninolytic fungi is an intracellular cytochrome P450 monooxygenase enzyme (CYP450)-stimulated reaction that forms arene oxide and water (Rathankumar et al., 2020). PAHs adsorbed by the fungal mycelium system are oxidized by the incorporation of a single oxygen molecule and transformed into unstable arene oxide, which is either non-enzymatically transformed into a phenol derivative or hydrated by epoxide hydrolase into trans-dihydrodiols (Dickson et al., 2019). The final phenol derivatives are subjected to enzymatic transformation and can act as substrates for subsequent methylation, sulfation, or conjugation with glucuronic acid, xylose, or glucose (Reyes-César et al., 2014).

Fungi are well-known organisms that can degrade lignocellulose, which is the most abundant renewable material available in nature, into desired monomeric products (Kumar and Chandra, 2020). Fungi produce lignolytic enzymes that break down polysaccharides, cellulose, hemicellulose, and lignin, which are natural aromatic polymers. Ligninolytic enzymes constitute a group of oxidoreductase enzymes that are highly specialized in the degradation of lignolytic biomass (Chukwuma et al., 2020). The structure of lignin macromolecules determines the properties of the degrading enzyme system, which is extracellular, nonspecific, and nonhydrolytic. White-rot fungi (WRF) are the most common fungal microbes and are potential lignolytic enzyme-producing microorganisms (Weng et al., 2021). They produce four main lignin-modifying enzymes, lignin peroxidase, manganese peroxidase, laccase, and versatile peroxidase, which are capable of completely mineralizing PHCs from contaminated soil.

White-rot fungi are potential microbes that are active against a wide range of organic compounds because they release extracellular lignin-modifying enzymes (LMEs) with low substrate specificity, allowing them to act on various molecules that are broadly similar to lignin (Dao et al., 2021). LMEs, such as lignin peroxidase (LiP), manganese peroxidase (MnP), various H₂O₂-producing enzymes, and laccase are potential enzymes for the degradation of lignin materials. However, these enzymatic activities may not be observed in lignolytic fungi. LiP, MnP, and versatile peroxidase (VP) have potential synergistic actions during lignin degradation (Falade et al., 2017). In addition to fungal oxidative enzymes, chemical reactions during lignin biodegradation produce fungal degraded lignin metabolites such as phenolics and other aromatic compounds, peptides, organic acids, and metal ions (Wang et al., 2018). It can be concluded that fungi, especially WRF, are potential microorganisms that can degrade a wide range of toxic and emerging contaminants, including PHC.

The mineralization of organic PCH to nontoxic inorganic compounds is heavily affected by the amount, composition, and inherent properties of the substrate. Evidence has shown that higher TPH concentrations in soil are toxic to microbes involved in degradation, and affect growth and remediation activity (Towell et al., 2011). However, extremely low TPHs provide marginal support for microbial growth and carbon supply as energy sources, leading to microbial decomposition of TPH inhibition in the soil. Moreover, the degree of PHC degradation is influenced by the chemical structure, composition, and bioavailability of PHC (Ramadass et al., 2018). Microbial decomposition of HCs can be ranked in the following order: linear HCs > branched HCs > low-molecular-weight alkyl aromatics > single aromatic rings > cyclic HCs > PAHs > asphaltenes (Koshlaf and Ball, 2017). Furthermore, biodegradability naturally depends on the composition of the contaminant, and less complex compounds are preferentially utilized by HC degraders (Shanmugaraju et al., 2013). For example, kerosene and saturated and/or aromatic crude oils, which consist of medium-chain alkenes, can be completely biodegraded. However, crude oil from heavy asphaltic-naphthenic acid may be degraded with low efficiency by several microbial species, but some novel strains, such as Pseudoalteromonas sp., can effectively degrade it under anaerobic conditions (Zan et al., 2022).

Temperature plays a paramount role in microbial remediation of organic contaminants (Varjani and Upasani, 2019). It affects both the chemistry and physiology of pollutants and the microbial flora (Chandra et al., 2013). Generally, temperature influences the microbial metabolism, microbial growth, solubility, and solid phases of the soil (Megharaj et al., 2011). For instance, higher temperatures enhance the water solubility of hydrophobic contaminants. In addition, higher temperatures decrease the viscosity of the contaminant, which enhances the diffusion of long-chain n-alkanes into the aqueous phase. Conversely, low temperatures cause increases in oil viscosity, reduced vaporization of noxious short-chain alkanes, and a decrease in water solubility, leading to a lag at the outset of microbial degradation (Ajona and Vasanthi, 2020). Although HC microbial degradation can occur over a range of temperatures, the speed of biological degradation decreases because of the limited enzymatic activity at decreasing temperatures (Fuentes et al., 2014). In addition, temperature affects the rate of degradation, and the physicochemical composition of oil leads to enriched hydrocarbon bioavailability in the microbial communities (Si-Zhong et al., 2009). Generally, the maximum bioremediation level occurs at 30°C–40°C in the soil environment, 20°C–30°C in freshwater systems, and 15°C–20°C in marine environments (Pithawala et al., 2012). Recently, researchers have investigated thermophilic microbes as alternatives to optimize the effect of temperature on the bioremediation of contaminants that can be operated at elevated temperatures (Aragaw et al., 2022). These microorganisms can adapt to stress and regulate their mechanisms under harsh conditions.

Nutrient accessibility also plays a significant role in microbial activity during the remediation of PH-contaminated soil. Typically, nutrients such as carbon, phosphorus, and nitrogen have a substantial impact on microbial growth for remediation of hydrocarbons (Dias et al., 2012). According to a review report by Varjani, PHCs are rich in carbon and energy sources but do not contain a significant amount of nitrogen and phosphorous needed to promote microbial growth (Varjani, 2017). However, the deficit in C-N-P-K ratios can be preserved by expanding CH₄ N₂ O, PO₄³⁻, N-P-K fertilizers, and NH₄⁺ with the most adequate 100:10:1 C: N:P ratio. Nitrate has been reported to be a nitrogen source for the growth and production of amphiphilic compounds (Toda and Itoh, 2012). As a result, petroleum-contaminated site treatment in the presence of nitrogen increases microbial cell growth and HC degradation rates. In addition, the presence of nitrogen in the culture decreases the PAH degrader growth lag phase of PAHs and maintains microbes at high pollutant degradation activity levels (Peng et al., 2008). However, it was revealed that an excess concentration of nitrogen in the soil causes inhibition, and maintaining its concentration (<1800 mg N₂/kg H₂O) provides optimal biological degradation of HC pollutants (Walworth et al., 2007). Likewise, the unbalanced high nutrient content of NPK inhibits biodegradation of HC contaminants (Boopathy, 2000). Furthermore, high quantities of NPK in the contaminated sites disrupted the C:N:P ratio, causing an oxygen deficit in the remediation sites. The soil type and organic matter in the soil significantly affected bioremediation. For example, the occurrence of high concentrations of soil organic matter in fine soils enhances microbial tumors and stimulates the mineralization of HCs (Scherr et al., 2007).

The soil type also plays a critical role in the success of bioremediation. A higher rate of hydrocarbon degradation occurred in the order of silky soil > sandy soil > clay soil because of the poor microbial content in the sand fraction and clay, which corresponded to a high C:N ratio and lower internal surface structure. For instance, more than 70% degradation of TPH was achieved in sandy soil and was relatively low in clay soil (Haghollahi et al., 2016). Clay soil develops a high surface area adhesive surface in water, resulting in a lower bioremediation rate due to the blockage of effective oxygen transmission through the soil. In addition, the robust adsorption of impurities on the surface of clay soil particles causes low degradation of HCS. In contrast, sandy soils have higher porosity than clays and permit superior oxygen transmission, allowing for the effective biodegradation of hydrocarbons. Moreover, the high porosity of sandy soils offers adequate space for microbial growth (Haghollahi et al., 2016). In conclusion, soil with a high porosity and oxygen transfer capacity has an improved HC bioremediation rate.

Oxygen is a pivotal element in the bioremediation of petroleum hydrocarbon-polluted soil, and the first biodegradation step begins with the presence of oxygen in PHC-polluted soil. Molecular oxygen is the most common electron acceptor element because most hydrocarbon substances are highly reduced substrates that require an electron acceptor. Providing a sufficient oxygen supply for the complete mineralization of PHC contaminants in soil is often challenging and expensive (Abbasian et al., 2015). In many aerobic bioremediation sites, oxygen is a crucial limiting factor that affects the complete decomposition of PHCs in the soil. Therefore, the addition of oxygen can significantly improve the remediation rate of microbial mineralization of PHCs and oxygenase-catalyzed remediation reactions (Waigi et al., 2015). It has been reported that aerobic microbial degradation is favorable for the participation of oxygenase and soil. The first steps in the catabolism of aliphatic (Peng et al., 2008), cyclic (Perry and Gibson, 1977), and aromatic (Elufisan et al., 2020) hydrocarbons by bacteria and fungi require molecular oxygen to be oxidized by oxygenase enzymes. However, the availability of oxygen in soil is dependent on a number of factors, such as the microbial oxygen consumption rate, type of soil, whether the soil is waterlogged, and the presence of utilizable substrates that can lead to oxygen depletion (Wang et al., 2021). The concentration of oxygen has been identified as a rate-limiting factor (Kumar et al., 2019). For example, soil usually requires 2 × 10⁶ m³ of water saturated with 10 mg/L O₂ for the effective oxidation of 10 m³ of hydrocarbons to carbon dioxide and water (Koshlaf and Ball, 2017). Therefore, dissolved oxygen in soil is crucial for the subsequent degradation of hydrocarbons and the active involvement of oxygenase enzymes.

The presence of active hydrocarbon-utilizing microorganisms is vital for bioremediation. Microbial groups such as bacteria, fungi, algae, protozoa, and viruses are the most widely used degraders of hydrocarbon contaminants (Ubani et al., 2016). These microbial strains can persist and use pollutants as sources of growth and metabolism in contaminated soils. Therefore, contaminated soils must contain a sufficient number of active hydrocarbon-utilizing microorganisms (Perelo, 2010). In contrast, the microbial inoculum size affects the breakdown of PHC pollutants. For instance, it has been reported that 106 CFU/g soil microbes are optimal for PHCs degradation (Towell et al., 2011).

Soil acidity and alkalinity are significant factors in petroleum hydrocarbon biodegradation. Unlike most aquatic ecosystems, soil pH can be extremely variable, ranging from 2.5 in mine spoils to 11.0 in alkaline deserts (Wang et al., 2021). Soil pH inhibits microorganisms and enzyme activity, whereas most microbes prefer neutral pH for growth and HC remediation (Kumar et al., 2019). Most microorganisms prefer neutral or low-alkaline conditions to decompose hydrocarbon compounds (Wang et al., 2016; Muangchinda et al., 2018). For instance, most heterotrophic bacteria and fungi prefer a pH near neutral, with fungi being more tolerant to acidic conditions (Al-Hawash et al., 2018a). Therefore, the ability of the microbial population to degrade hydrocarbons is expected to be negatively affected by increased pH, as has been observed in some soils. Hence, the optimum microbial growth rate and HC biodegradation can be maintained at pH values between 6 and 9 with inadequate nutrient and oxygen supplies (Das and Chandran, 2011). Additionally, PHCs have ecotoxic effects on bacterial activity, plant species, and earthworms, resulting in reduced biodegradation rates (Tang et al., 2011).

Soil salinity also affects plants and microorganisms. Higher salinity may decrease the growth of microbes in water and soil. Additionally, some microbial communities are more prone to salinity. Increased salt concentration reduces microbial growth by inhibiting enzyme secretion due to the high osmotic pressure of the surrounding environment (Qin et al., 2012; He et al., 2017), reducing available oxygen to microorganisms (Liang et al., 2017) and hydrocarbon compound solubility (Truskewycz et al., 2019), and also causes cell membrane lysis (Muangchinda et al., 2018). Typically, extreme salt concentrations inhibit bacterial growth and cause bacterial cell dehydration, leading to cell death. Furthermore, high salinity levels decrease the HC metabolism rate as a source of carbon and energy. Consequently, high salt concentrations deactivate HC degraders and soil decontamination becomes ineffective (Rath et al., 2016). Several investigations have shown that lower salinities increase mineralization of HCs, but extremely low salinities significantly reduce biodegradation (Qin et al., 2012). Table 3 shows the factors affecting bioremediation of PHC-polluted soils.

Petroleum hydrocarbon biodegradation metabolic pathways and routes are poorly understood, and inefficient transformation results in the accumulation of toxic intermediate substances that can resist further degradation by existing pathways. Hence, appropriate inoculum selection, optimized inoculation time, reaction conditions (aerobic, anoxic, and anaerobic), and physicochemical parameters are critical for the complete and effective bioremediation of petroleum hydrocarbon contaminants. Furthermore, recombinant DNA approaches have been introduced in the microbial bioremediation of contaminants to evolve the desired gene manipulation of native microorganisms and their degradation pathways to enhance their bioremediation potential by obtaining novel contaminant-degrading genes and enzymes. Although several approaches have been developed in the field of bioremediation of petroleum hydrocarbons, much remains to be explored. In this regard, the strengths, weaknesses, opportunities, and threats in the field of PHC remediation are crucial for developing a comprehensive understanding of bioremediation prospects.

Strengths:

• PHC-contaminated sites can be cleaned in a sustainable and eco-friendly manner by utilizing natural microbial processes.

Weaknesses:

• Microbial degradation of contaminants, including PHCs, is time consuming.

Opportunities:

• Continuing progress in biotechnology can lead to the creation of more efficient and robust microorganisms for PHCs bioremediation.

Threats:

• Traditional chemical remediation techniques may still be preferred because of their apparent speed and instant results, posing a threat to the widespread adoption of bioremediation.

The use of petroleum and petroleum products as energy sources contributes to oil spills in the environment, presenting a challenge for the remediation of polluted sites. Consequently, petroleum hydrocarbon contamination poses significant environmental pollution, and bioremediation has emerged as a promising and sustainable approach for mitigating its impact on soil ecosystems. This review comprehensively evaluates the current state of knowledge to contribute to a better understanding of the complexities involved in bioremediation processes and highlights potential avenues for future research and applications in the restoration of PHC-contaminated soil. From the evaluation of the aforementioned discussion of bioremediation efforts, the following conclusions and recommendations were drawn.

• Microbial consortia, such as bacteria and fungi, and the degradation mechanisms employed by them showcase the diverse and efficient nature of biological processes in breaking down complex hydrocarbons.