A review on biosurfactant producing bacteria for remediation of petroleum contaminated soils

Abstract

The discharge of potentially toxic petroleum hydrocarbons into the environment has been a matter of concern, as these organic pollutants accumulate in many ecosystems due to their hydrophobicity and low bioavailability. Petroleum hydrocarbons are neurotoxic and carcinogenic organic pollutants, extremely harmful to human and environmental health. Traditional treatment methods for removing hydrocarbons from polluted areas, including various mechanical and chemical strategies, are ineffective and costly. However, many indigenous microorganisms in soil and water can utilise hydrocarbon compounds as sources of carbon and energy and hence, can be employed to degrade hydrocarbon contaminants. Therefore, bioremediation using bacteria that degrade petroleum hydrocarbons is commonly viewed as an environmentally acceptable and effective method. The efficacy of bioremediation can be boosted further by using potential biosurfactant-producing microorganisms, as biosurfactants reduce surface tension, promote emulsification and micelle formation, making hydrocarbons bio-available for microbial breakdown. Further, introducing nanoparticles can improve the solubility of hydrophobic hydrocarbons as well as microbial synthesis of biosurfactants, hence establishing a favourable environment for microbial breakdown of these chemicals. The review provides insights into the role of microbes in the bioremediation of soils contaminated with petroleum hydrocarbons and emphasises the significance of biosurfactants and potential biosurfactant-producing bacteria. The review partly focusses on how nanotechnology is being employed in different critical bioremediation processes.

Introduction

Environmental contamination caused by petroleum products such as diesel, gasoline and crude oil has gained ecological attention (Jimoh and Lin 2019). Such pollution is caused by various activities, including industrial runoffs, effluent release, offshore and onshore petroleum industry operations and accidental spills (Yuniati 2018). Leaks and spills are expected during the exploration, production, refining, transportation, processing, as well as storage of petroleum and petroleum products (Liu et al. 2016a; Kvenvolden and Cooper 2003). Human activity-induced unintentional or deliberate release of hydrocarbons into the environment is one of the major sources of soil and water pollution (Das and Chandran 2011). Contamination of surface soil, groundwater and the ocean occurs due to the release of petroleum hydrocarbon by spills along with leaks from underground tanks, steamers, unplugging of oil wells and abandoned oil refinery sites (Souza et al. 2014; Janbandhu and Fulekar 2011; Saeki et al. 2009; Prince et al. 2013). Crude oil spills often occur during the transport of oil and significant volumes of residues and wastewater containing oil are produced during oil extraction and processing (Deng et al. 2014). Total petroleum hydrocarbon (TPH) concentrations in polluted soils around the world are reported to be in the critical range of 1.17–236.7 g per kilogram of soil (Waychal et al. 2022). However, annually natural crude oil seepage is estimated to be 600,000 metric tonnes, with a 200,000 metric tonne per annum range of uncertainty (Kvenvolden and Cooper 2003).

Petroleum waste is a mixture of water, emulsion, solids, and liquid crude oil (Ite et al. 2013). Further, saturates, aromatics, asphaltenes (phenols, fatty acids, ketones, esters, porphyrins, etc.) and resins (pyridines, quinolines, carbazoles, sulfoxides, amides) are the four major groups of petroleum hydrocarbons (Colwell et al. 1977; Yuniati 2018). Hydrocarbons accumulate in animal and plant tissue causing mutations and mortality and thus are responsible for considerable damage to the local system. Due to their adverse impact on human health and the environment, US Environmental Protection Agency (1986) categorized these chemicals as priority environmental pollutants. These organics hardly break down in the soil, posing a hazard to animal and plant species, causing harm to human health and ecosystem (Khan et al. 2016). Oil permeation into the soil makes the degradation process extremely difficult. The hydrocarbons are highly hydrophobic, have less water solubility, and get attached to soil particles, resulting in decreased bioavailability to microorganisms, thus limiting the mass transfer rate in biodegradation (Iwabuchi et al. 2002). Therefore, pollutants must be transferred into a bulk phase to improve the bioavailability of the pollutants in soil (Aqib Hassan et al. 2017).

For the remediation of hazardous petroleum pollutants, a range of physical (burning, land disposal, landfilling, electrokinesis), chemical (application of nanoparticles, addition of oxygen, use of surfactant) and biological (use of plants and plant debris, microbes) approaches are employed (Cristovao et al. 2016). However, the cost of excavating and transporting large volumes of contaminated materials for ex-situ treatment makes decontamination process expensive. Furthermore, these methods are costly and correspondingly result in partial pollutant breakdown. Therefore, green technologies for pollutant clean-up by biological means are used for bioremediation of petroleum polluted site(s) (Rahman et al. 2003; Varjani et al. 2015).

Employing microorganisms to detoxify or eliminate pollutants due to their wide metabolic capacities is a developing strategy for the removal as well as degradation of petroleum hydrocarbons and various other environmental contaminants (Medina-Bellver et al. 2005). Moreover, bioremediation, or biodegradation by natural populations of microorganisms, is thought to be non-invasive and cost-effective (April et al. 2000). Biodegradation of xenobiotic pollutants (such as oil) by natural populations of microorganisms is more ecologically friendly, cost-effective and efficient than conventional physico-chemical treatments (Das and Mukherjee 2007; Zhang et al. 2012; April et al. 2000; Cappello et al. 2007). Although short-chain petroleum hydrocarbons are degraded by many bacteria, however, microbes that can digest long-chain petroleum hydrocarbons (LCPHs) and complex polycyclic aromatic hydrocarbons (PAHs) are crucial for oil pollution clean-up (Kenzo et al. 2008). Bioaugmentation, involving addition of known oil degrading bacteria to supplement the existing microbial population, and biostimulation, involving stimulation of the growth of indigenous oil degraders by the addition of nutrients or other growth-limiting co-substrates, are the two main approaches to bioremediation of oil contaminated soils (Das and Chandran 2011). In recent decades, bioremediation has gained significant appreciation due to its natural approach, likely to be incorporated as a process of integrated treatment methods along with physico-chemical method for pollutant removal (Chrispim and Nolasco 2017; Diplock et al. 2009).

Inoculation of hydrocarbon degraders to polluted sites is one of the very long practiced remediation strategies; however, the effectiveness of microbial biodegradation is often limited by the low bioavailability of insoluble hydrocarbons compounds to microorganisms (Barathi and Vasudevan 2001; Verma et al. 2006; Nitschke and Pastore 2006). Therefore, employing biosurfactant-producing bacterial strains can be a significant approach in the realm of petroleum hydrocarbon bioremediation (Pacwa-Płociniczak et al. 2011).

Biosurfactants, first discovered in late 1960s, are extracellular amphiphilic compounds formed on living surfaces, primarily microbial cell surfaces, or discharged outside the cell. They are made up of both hydrophobic and hydrophilic moieties that remain distributed at the interfaces between liquid phases with varying degrees of polarity (oil/water), lowering surface and interfacial tension, thereby increasing the bioavailability & removal percentage of the organic pollutants like petroleum hydrocarbons by pseudosolubilization together with emulsification (Mondal et al. 2019). Furthermore, biosurfactants stimulate bacteria to putrefy toxins, facilitating in-situ bioremediation as well as solubilization and desorption of soil pollutants. Microorganism produces biosurfactants that act as emulsifiers, enabling them to utilize petroleum hydrocarbons as an energy source, thereby mineralizing or converting hydrocarbons to harmless alternatives (Priya and Usharani 2009).

Besides possessing numerous appealing characteristics such as minor or no toxicity, ease of production, renewable resources, natural degradability and high activity, biosurfactants play an insignificant role commercially as compared to their unnatural counterparts, i.e. synthetic surfactants (Thavasi et al. 2011; Shubina et al. 2016). While developing a plan for more significant hydrocarbon degradation aided by biosurfactants, various factors must be considered (Mukherjee et al. 2006). Biosurfactant generation and efficacy are influenced by temperature, limiting nutrients (such as N, Fe, Mg, etc.), culture agitation speed, pH and salinity (Amézcua-Vega et al. 2007; Oliveira et. al. 2009a; Wei et al. 2003; Zeraik and Nitschke 2010; Silva et al. 2010; Abouseoud et al. 2010; Bai et al. 1998).

Selecting appropriate biodegraders, availability of nutrients, suitable viscosity and adequate concentration of petroleum waste are some of the limitations of biological treatments such as bioremediation (Zahed et al. 2010). Challenges in bioremediation confirmed that thorough remediation of petroleum-contaminated soils requires an additional accelerator with non-toxic properties together with a broad surface area to speed up the bioremediation process (Alabresm et al. 2018). Nanoparticle-assisted bioremediation could be a viable alternative to the limitations of the bioremediation process. Nanoparticles have been successfully employed commercially to remediate contaminated soils due to their eco-friendly behaviosur, environmental feasibility, enhanced adsorption capacity, lowest reactivity, the release of harmless ions and massive surface area for bacteria (Adeleye et al. 2016). Moreover, NPs assist in bioremediation either by accelerating growth of microbes, immobilising microbial cells, or inducing the synthesis of microbial enzymes. Furthermore, nanoparticles stimulate the formation of biosurfactants and solubility of hydrophobic organic hydrocarbons, thereby allowing microbes to readily degrade these organic contaminants. Unrestricted release of nanoparticles into the environment, on the other hand, will have a negative impact on abiotic and biotic constituents along with human health. As a result, before these nanoparticles are used on a large scale, a deeper understanding of their environmental fate must be investigated. The present review investigates the role of microorganisms in remediation of petroleum contaminated soils in addition to the importance of biosurfactant and nanoparticles in assisting bioremediation process.

Direct or indirect effects of wastewater or substrate pollution by PAH on plants

Total Petroleum hydrocarbons (TPH) are one of the most prevalent classes of persistent organic pollutants while soil contamination with petroleum hydrocarbons is amongst the most widespread environmental problems worldwide (Huang et al. 2005; US EPA 2000). Moreover, many species of plant are vulnerable to petroleum pollution (Huang et al. 2004). For instance, in a phytoremediation investigation, ryegrass biomass was shown to have decreased by 96% after 30 days of growth on soil containing petroleum hydrocarbons at the rate 25 g per kg (Tesar et al. 2002). Likewise, the wastewater produced by petrochemical activities contains a variety of harmful contaminants, including heavy metals and compounds like ammonia, diammonium phosphate, phosphoric acid, sulphuric acid, etc. (Alavi and Amir-Heidari 2013; Tehrani et al. 2012; Malmasi et al. 2010). Therefore, using such water to irrigate crops resulted in decreased plant yield and biomass because contamination from heavy metals inhibits the assimilation of nutrients necessary for plant growth such as magnesium (Mg), potassium (K), phosphorus (P), sulphur (S) iron (Fe), Calcium (Ca), zinc (Zn), manganese (Mn) and copper (Cu), thus disrupting their translocation from roots to other plant parts (Benavides et al. 2005; Lefevre et al. 2014; Shanker et al. 2005; Banon et al. 2011; Sharma and Agarwal 2005; Lyu et al. 2016). Since some minerals are essential for plant growth, and when their concentrations fall below optimal levels, the health of the plant is affected along with the decrease in overall yield (Imtiaz et al. 2015). Hajihashemi et al. (2020) studied the effects of wastewater released from petrochemical industries on two wheat cultivars, i.e.. Chamran and Behrang. They reported that the toxic levels of mineral elements in the wastewater resulted in a significant decline in the potassium, silicon and zinc content of leaves.

Furthermore, one of the most prevalent and harmful pollutants are polycyclic aromatic hydrocarbons (PAHs), which are compounds with at least two benzene rings (Fang et al. 2006). Owing to their lipophilic nature, PAHs predominantly affect the lipids in cell membranes, altering their permeability, consequently upsetting the ionic balance and barrier characteristics (Kreslavski et al. 2017). It has been demonstrated that the presence of certain PAHs such as naphthalene, phenanthrene and fluoranthene, impairs the operation of the photosynthetic apparatus (PA), inhibits the activity of photosystem 2 (PS-2) and slows down photosynthetic electron transport, thereby adversely affecting various photosynthetic activities (Kummerova et al. 2001, 2006; Tomar and Jajoo 2013, 2014; Tomar et al. 2015). According to Lankin et al. (2014), the photochemical activity of PS-2 in detached pea leaves was decreased by the short-term exposure to naphthalene. Besides, it has been hypothesised that PAHs have an impact on the amount of PS-2 QB-non-reducing centres as well as on the stability of the light-harvesting antenna complexes (Kreslavski et al. 2017).

In addition to this, it was reported that irrigation of wheat with untreated effluent from petrochemical industry affects photosynthetic properties such as net photosynthesis, intercellular carbon dioxide, photosynthetic pigments and water use efficiency (Hajihashemi et al. 2020). Thereby, subsequently reducing the concentration of carbohydrates and nutrients, followed by reduction in leaf area, plant height and grain productivity. Furthermore, it was noticed that increase in the concentration of the wastewater for irrigation reduces diameter of root and thickness of leaf, which account for the thinned mesophyll and reduced number xylem and phloem channels in the parenchyma of root cortex (Hajihashemi et al. 2020).

Water and nutrient uptake by plants are, moreover, negatively impacted by PAH contamination, which has an influence on biomass yield (Reilley et al. 1996; Cheema et al. 2010; Oguntimehin et al. 2010). On investigating the impact of phenanthrene, fluoranthene, as well as benzo[a]pyrene contamination on growth and biomass of three plants, i.e. Medicago sativa, Lolium perenne and Festuca arundinacea, Afegbua and Batty (2018) reported an increase in shoot biomass yield of 110% for M. sativa, 8% decrease in root biomass yield for L. perenne and 12% decrease in F. arundinacea shoot biomass yield. An increase in the biomass and shoot-to-root ratio of Medicago sativa indicates resistance for PAH pollution and a beneficial impact on plant growth (Hall et al. 2011). However, a decline in the shoot-to-root ratio for L. perenne and F. arundinacea suggests that the PAH treatments had phytotoxic effects, adversely affecting plant growth, thus causing senescence (Kechavarzi et al. 2007; Cheema et al. 2010). However, Cheema et al. (2010) reported a 35% decrease in M. sativa biomass in soils spiked with phenanthrene (200 mg kg⁻¹) and pyrene (199 mg kg⁻¹). Phytotoxicity depends on a plant's ability to withstand stress and on the ability of the native soil bacteria to degrade that toxic compound (Kechavarzi et al. 2007).

Microbes mediated degradation of petroleum hydrocarbons

Petroleum hydrocarbon biodegradation is a complicated process driven by the chemical structure, i.e., the amount and type of hydrocarbons present. Some bacteria, for instance, Achromobacter xylosoxidans DN002 (Ma et al. 2015), Pseudomonas spp. (Venkateswaran et al. 1995), Bacillus Licheniformis (Eskandari et al. 2017) can break down aromatic or resin fractions of hydrocarbons, whereas others such as Dietzia spp. (Wang et al. 2011a, b), Oleispira antarctica (Yakimov et al. 2003), Geobacillus thermodenitrifican (Abbasian et al. 2015) can degrade specific alkanes (Xu et al. 2018). Indigenous bacteria ultimately degrade or metabolize petroleum hydrocarbons because they utilize hydrocarbons as energy and carbon sources for growth and reproduction and relieve physiological stress caused by petroleum hydrocarbons in the microbial bulk environment (Hazen et al. 2010; Kleindienst et al. 2015). The resistance of hydrocarbons to microbial action varies and follows general ranking as follows: asphaltenes > polyaromatics > cyclic alkanes > monoaromatics > low molecular weight alkyl > aromatics > branched alkanes > linear alkanes (Varjani 2017). Different bacterial strains have been identified and employed as efficient biodegraders to treat soil contaminated with petroleum hydrocarbons. Microorganisms carrying out bioremediation are either indigenous (native) or exogenous. The activity of native microorganisms can be enhanced by adding nutrients, while the addition of exogenous bacteria to the contaminated site promotes bioremediation. Because native bacteria take a long time to domesticate, they have poor growth rates and metabolic activity, making decontamination time-consuming and ineffectual. Further, alkanes with shorter and longer carbon chains (C10 and C20–C40) and polycyclic aromatic hydrocarbons (PAHs) are difficult to degrade. Therefore, not all crude oil components can be removed using hydrocarbon-degrading bacteria in the oil-contaminated site.

However, petroleum hydrocarbon molecules adhere to soil particles and are difficult to degrade; hence, the limited availability of oil pollutants to microorganisms is one of the significant factors limiting biodegradation in the environment (Barathi and Vasudevan 2001). Therefore, the most crucial phase in the biodegradation of petroleum hydrocarbons, according to Shi et al. (2019) is developing the interface between substrates and petroleum-degrading bacteria. This interface can currently be developed in three ways, viz. (1) microbial cells dissolve petroleum hydrocarbons in an aqueous solution and absorb them, (2) large hydrocarbon molecules are immediately exposed to microbial cells for absorption, and (3) small pseudo-soluble, quasi-soluble, or encapsulated hydrocarbon particles interact with microbial cells for absorption. The hydrophobicity of bacterial cells, on the other hand, affects their adherence to petroleum hydrocarbons. Therefore, in order to improve the efficacy of the bioremediation process, high hydrophobicity of bacterial cells is required, which in turn needs the application of biosurfactants for efficient interaction between bacteria and petroleum hydrocarbons (Lin et al. 2005; Pan et al. 2018; Zhang et al. 2012).

Patil et al. (2012), studied diesel oil degradation in oil-polluted soil & identified Bacillus spp., Acinetobacter spp., Micrococci spp., Pseudomonas spp. and Streptomyces spp. as potential hydrocarbon-degrading microbes. Liu et al. (2016b), isolated high salinity, alkalinity and temperature-tolerant efficient hydrocarbon-degrading strain Bacillus licheniformis from heavily oil-contaminated soil around the Dagang Oilfield Tianjin, China. The reported strain degrades both short-chain alkanes and long-chain alkanes with their more complex structures and secretes emplastic at high temperatures, which could be applied as a surfactant to augment the emulsifying effect. In addition to this, Verma et al. (2006), assessed oily sludge degradation capabilities of three bacterial isolates from a contaminated site in Ankleshwar, India and based on gravimetric analysis, reported that Bacillus spp., Acinetobacter spp. and Pseudomonas spp., respectively, degrade approx. 59%, 37% and 35% of the oily sludge in 5 days at 30 °C. Further, the capillary gas chromatographic analysis confirmed that after 5 days, the Bacillus spp. was able to metabolize oily sludge components of chain length C12–C30 and aromatics more efficiently than the other identified two strains. Jiji and Prabakaran (2020), isolated six native oil-degrading bacterial species from a petroleum-contaminated soil at Changanassery, Kottayam district in Kerala, India, identified as Bacillus cereus, Pseudomonas aeruginosa, Bacillus subtilis, Pseudomonas spp, Bacillus spp. and Staphylococcus aureus. Das and Mukherjee (2007) reported that thermophilic strains of Bacillus subtilis and Pseudomonas aeruginosa indigenous to North-East India were efficient in the biodegradation of crude petroleum oil. Likewise, the most critical hydrocarbon-degrading bacterial genera in contaminated soils, according to published literature, include Achromobacter, Arthrobacter, Bacillus, Nocardia, Nocardioides, Pseudomonas, Rhodococcus, Sphingomonas, Variovorax and other unculturable bacterial clones (Jiji and Prabakaran 2020).

Few bacterial species can degrade a wide range of petroleum hydrocarbons; for instance, according to the study conducted by Wang et al. (2011a, b), Dietzia spp. DQ12-45-1b utilizes n-alkanes (C6–C40) and other compounds as the sole carbon sources, and Ma et al. (2015) reported that Achromobacter xylosoxidans DN002 degrades a variety of monoaromatic and polyaromatic hydrocarbons. However, most bacterial species can metabolize only a limited number of petroleum hydrocarbon components, leaving others inaccessible (Xu et al. 2018; Chaerun et al. 2004; Varjani 2017). This is because different native bacterial species possess specific catalytic enzymes and thereby, to achieve the efficient bioremediation of petroleum hydrocarbon-contaminated soils, the joint action of different functional bacteria is required (Dombrowski et al. 2016). In supporting the above statements, Varjani et al. (2015) reported that halotolerant Hydrocarbon Utilizing Bacterial Consortium (HUBC) comprises Ochrobactrum spp., Stenotrophomonas maltophilia, and Pseudomonas aeruginosa efficiently degrades 3% v/v crude oil at the rate of 83.49%. Szulc et al. (2014) conducted a field study and confirmed biodegradation efficiency of 89% in a 365-day treatment of diesel oil-contaminated soil by an artificial consortium comprising Alcaligenes xylosoxidans, Aeromonas hydrophila, Pseudomonas fluorescens, Gordonia spp., Rhodococcus equi, Pseudomonas putida, Stenotrophomonas maltophilia and Xanthomonas spp. Because bacterial consortium possess variety of catabolic genes, the synergistic effects of these genes are advantageous in accomplishing pollutant purification (Gurav et al. 2017).

Therefore, the presence of variety of catabolic genes in a bacterial consortium, and the synergistic effects of these genes are beneficial for achieving the purification of pollutants (Gurav et al. 2017). A bacterial consortium comprising of Bacillus strain, two Mycobacterium strains, Novosphingobium and Ochrobactrum showed synergistic pyrene degradation due to the following aspects: (1) By synthesizing biosurfactant, the Bacillus strain increased pyrene bioavailability, (2) Pyrene breakdown process was started by two Mycobacterium strains and (3) the pyrene intermediates were efficiently destroyed by Novosphingobium and Ochrobactrum (Wanapaisan et al. 2018). However, the construction of a minimal functioning bacterial consortium or genetically engineered bacteria for bioremediation of petroleum oil has been required due to the complexity of the petroleum components (Dvorak et al. 2017). Moreover, community stability and the safety of the modified bacteria are two additional issues that must be addressed (Xu et al. 2018). Hence, the above reports suggest that an application of bacterial consortia could be a feasible and reasonable approach for efficiently removing petroleum hydrocarbons from contaminated environments.

Biodegradation mechanism of petroleum hydrocarbons

A significant portion of organic contaminants degrades most quickly and entirely under aerobic environment. The first step in aerobic catabolism appears to be adding one or two hydroxyl groups to the hydrocarbon skeleton (Chandra et al. 2013). The activation and inclusion of oxygen catalyzed by oxygenases and peroxidases are the principal enzymatic reaction during the initial intracellular degradation of organic contaminants; thus it is an oxidative process (Fritsche and Hofrichter 2000; Rojo 2009). Monooxygenases catalyze the introduction of one oxygen atom into the hydrocarbon, and dioxygenases catalyze the inclusion of two hydroxyl groups (Fuentes et al. 2014). Organic pollutants are converted into the metabolites of central intermediary metabolism, such as the tricarboxylic acid cycle, through peripheral degradation mechanisms. The primary precursor metabolites, such as acetyl-CoA, succinate and pyruvate, are utilized in cell biomass synthesis and gluconeogenesis, producing sugars needed for other biosynthetic reactions involved in growth and development. Hence, the microorganism can survive in a nutrient-limited environment by oxidizing these substrates. However, during anaerobic degradation, sulfate and nitrate act as terminal electron acceptors and breakdown is accomplished by coupling CO₂ or fumarate to hydrocarbons (Callaghan et al. 2012; So et al. 2003). However, when compared to aerobic microbial catabolism, the anaerobic breakdown of petroleum hydrocarbons happens slower (Wentzel et al. 2007). Figure 1 highlights the basic concept of aerobic hydrocarbon decomposition (Das and Chandran 2011).

Fig. 1.

Basic principle of aerobic hydrocarbon breakdown by microorganisms. Adapted and redrawn from: Das and Chandra (2011). NH4+ Ammonium ion, PO43- Orthophosphate, SO42- Sulfate, Ferric iron; CO2 carbon dioxide, H2O water, O2 Oxygen, TCA Tricarboxylic acid cycle

Microbial cell adhesion to substrates and biosurfactant synthesis are the other mechanisms involved in the biodegradation of organic contaminants (Das and Chandran 2011). Although the absorption mechanism involving cell attachment to an oil droplet is still unexplained, biosurfactant synthesis has been extensively studied (Chandra et al. 2013). According to Perfumo et al. (2006), a novel hydrocarbon-degrading thermophillic bacteria Pseudomonas aeruginosa strain APO2-1 produces rhamnolipid while growing at a temperature of 45 °C and using water insoluble and soluble hydrocarbon as substrate. Similar to this, when grown on crude oil, Aeromonas spp. isolated from tropical estuary water, produces biosurfactant that can emulsify a variety of hydrocarbons, with most effective substrate as diesel (emulsification index after 24 h, i.e. E₂₄ = 65) and least effective substrate as hexane (E₂₄ = 22) as the least effective. Moreover, following purification, it was discovered that the biosurfactant synthesized by Aeromonas spp. contained an unknown lipid and around 38% of carbohydrates and showed absence of proteins (Ilori et al. 2005). It has been reported that a strain of Pseudomonas aeruginosa produces a mono- and di-rhamnolipid type biosurfactant (Patel et al. 2015). Hence, by lowering surface tension and generating micelles, biosurfactants can function as emulsifying agents and enable the encapsulation of oil micro-droplets in a hydrophobic microbial cell surface, which is then taken inside by microbial cell and then destroyed (Chandra et al. 2013).

Enzymes involved in degradation of petroleum hydrocarbons

Enzymes act as biocatalysts and speed up the reaction rate by lowering the activation energy, allowing faster degradation (Patel et al. 2020). Microbes or their extracted biological components, such as enzymes, are the principal components of bacterial degradation of petroleum hydrocarbons (Wasmund et al. 2009; Varjani 2017). Due to the vast substrate specificity, enzymes play a significant role in the biodegradation of phenols, PAHs and other petroleum-based hazardous chemicals (Patel et al. 2016, 2019). For instance, the degradation of alkanes is catalyzed by alkane-1-monooxygenase, alcohol dehydrogenase, methane monooxygenase, cyclohexanol-dehydrogenase and cyclohexanone 1, 2-monooxygenase. However, naphthalene degradation is carried out by naphthalene 1, 2-dioxygenase ferredoxin reductase component, cis-2,3-dihydrobiphenyl-2,3-diol dehydrogenase and salicylaldehyde dehydrogenase, while benzene dioxygenase, toluene dioxygenase and ethylbenzene dioxygenase catalyse degradation of other petroleum contaminants (Bacosa et al. 2018). Cytochrome P450 alkane hydroxylases enzymes involved in the microbial breakdown of chlorinated hydrocarbons, oil, fuel additives and various other chemical compounds, belongs to the superfamily of heme-thiolate monooxygenase (Van Beilen and Funhoff 2007). To initiate biodegradation, enzyme systems must incorporate oxygen into the substrate, depending on the chain length. Eukaryotes of higher-order possess different P450 families, each consisting of a considerable number of unique P450 forms that work together to convert a specific substrate metabolically. However, in prokaryotic microorganisms, such P450 diversity is detected in only a few species (Zimmer et al. 1996). The existence of diverse microsomal Cytochrome P450 forms enables certain yeast species such as Candida maltose, Candida tropicalis and Candida apicola to utilize n-alkanes and other aliphatic hydrocarbons as their primary carbon and energy source (Scheuer et al. 1998).

Integral membrane di-iron alkane hydroxylases (e.g., alkB), Cytochrome P450 enzymes, membrane-bound copper-containing methane monooxygenases, and soluble di-iron methane monooxygenases are the examples of alkane oxygenase systems in prokaryotes and eukaryotes that are actively engaged in the degradation of alkanes under aerobic conditions. (Van Beilen and Funhoff 2005). Besides oxygenases, several other enzymes have been utilized in the bioremediation of petroleum toxicants, including oxidoreductases, laccase and peroxidase (Okino-Delgado et al. 2019; Patel et al. 2017, 2018). Different enzymes produced by potential microorganisms to break down various contaminants are listed in Table 1.

Table 1.

Enzymes derived from various microbes to degrade petroleum hydrocarbons

Enzymes	Microorganisms	Substrates	References
Particulate methane monooxygenases	Methylobacter, Methylococcus, Methylocystis spp.	C1–C5 (halogenated) alkanes and cycloalkanes	McDonald et al. (2006)
Dioxygenases	Acinetobacter spp.	C10–C30 alkanes	Maeng et al. (1996)
Soluble methane monooxygenases	Methylococcus, Methylosinus,Methylocystis Methylomonas, Methylocella	C1–C8 alkanes alkenes and cycloalkanes	McDonald et al. (2006)
Bacterial P450 oxygenase system	Acinetobacter, Caulobacter, Mycobacterium	C5–C16 alkanes, cycloalkanes	Van Beilen et al. (2006)
AlkB related alkane hydroxylases	Pseudomonas Burkholderia, Rhodococcus, Mycobacterium	C5–C16 alkanes, fatty acids, alkyl benzenes, cycloalkanes	Van Beilen et al. (2003)
Eukaryotic P450	Candida maltose, Candida tropicalis, Yarrowia lipolytica	C10–C16 alkanes, fatty acids	Iida et al. (2000)
Lignin Peroxidases, Manganese Lignin peroxidases, Manganese peroxidases and laccases	Trametes versicolor, Pleurotus ostreatus and Phanerochaete chrysosporium	Pyrene and anthracene	Novotny et al. (2004)
Laccase, Cytochrome- P450	Pycnoporus sanguineus	Anthracene, Pyrene	Zhang et al. (2015)
Lipase, catalase & oxidoreductase	Shewanalla chilikensis, Bacillus firmus, Halomonas hamiltonii	Total Petroleum Hydrocarbon	Suganthi et al. (2018)
Laccase, lignin peroxidase, manganese peroxidase	Ganoderma lucidum	Phenanthrene, Pyrene	Agrawal et al. (2018)
Laccase	Trametes versicolor	Anthracene	Apriceno et al. (2017)

Adapted from: Chandra et al. (2013) and Das and Chandra (2011)

Microbial uptake and trans-membrane transport of petroleum hydrocarbons

The hydrophobic nature of bacterial cell membrane limits the availability of hydrocarbons for bacterial cell uptake, thus, making biodegradation more difficult. Henceforth, most substrates that facilitate microbial growth involve adherence to a bacterial cell before the catabolic machinery of the cell can access them. Intracellular and extracellular degradation are two ways involved in microbial degradation of petroleum contaminants (Wang et al. 2020). The intracellular localization of hydrocarbon-degrading enzymes indicates that the whole process of degradation can be divided into the following three steps:

1. Adsorption and uptake of hydrocarbon compounds by microorganisms.

2. Entry of the hydrocarbons adsorbed on the cell surface into the cell membrane via trans-membrane transport.

3. Degradation of hydrocarbons by degradation enzymes localized on the cell membrane (Hua and Wang 2014; Eguchi et al. 1992).

Hydrocarbon uptake by microorganisms

Uptake of hydrocarbons can occur through the following three main routes: direct adhesion, emulsification and pseudo-solubilization and interfacial uptake.

Dissolved hydrocarbon uptake in the aqueous phase, i.e., direct adhesion

Due to poor solubility (10^–10–10^–5 gramme per litre; g/L) of long-chain hydrocarbons or other high molecular weight substrates, direct adhesion is applicable for only short-chain hydrocarbons and water-soluble aromatics (Ratledge 1992). Hence, it is known as the "single diffusion–dissolution uptake" model (Wang et al. 2020). Since microorganisms utilize dissolved hydrocarbon as carbon and energy sources for growth, they directly adsorb hydrocarbon dissolved in water. Therefore, the solubility of hydrocarbons influences the hydrocarbon adsorption and uptake efficiency of microbe. In general, microorganisms grow successfully when their growth rate is slower than the rate at which hydrocarbons solubilize because enough carbon and energy source will be present in the environment for their growth (Rosenberg et al. 1980; Hua and Wang 2014). Conversely, a microorganism's biodegradation capability is limited if the microorganisms grow faster than at which hydrocarbons dissolve (Hua and Wang 2014). Furthermore, extracellular polymeric substances (EPS) play a crucial role in direct adsorption because bacterial EPS plays a pivotal role in producing biofilms on other surfaces (Pan et al. 2012).

Uptake of pseudosolubilized hydrocarbon droplets: emulsification and pseudosolubilization

The hydrocarbon uptake approach involves synthesizing emulsifying agents such as biosurfactants to form micro-droplets of hydrocarbon. It has been proved that biosurfactants possess amphiphilic properties and emulsify hydrocarbon leading to improved water solubility and reduced surface tension (Nguyen et al. 2016; Ronning et al. 2015; Helfrich et al. 2015). In addition, biosurfactant accelerates the displacement of oily substances from soil particles, assisting the transport and translocation of the insoluble substrates across cell membranes and helping dissociation of bacteria from the oil droplets after the utilizable hydrocarbon has been depleted (Abalos et al. 2004; de Carvalho et al. 2005; Hua et al. 2004; Prabhu and Phale 2003).

Direct contact with large hydrocarbon drops: interfacial uptake

Microbial cells directly adhere to liquid hydrocarbon drops that are significantly larger than the cells in this approach of hydrocarbon uptake. Microorganisms growing in the fatty acids rather than the water frequently form an agglomeration with hydrocarbons, bringing the microbe cells and hydrocarbons into intimate contact. Goswami and Singh (1991) reported that the Pseudomonas strain transports the hexadecane directly into the cells without an emulsifier or extracellular surfactant by adhering to the substrate surface. Furthermore, Coimbra et al. (2009), discovered that yeast cells have high or intermediate interfacial tension and hydrophobicity and low surface tension, indicating that the cells employ two insoluble substrate uptake techniques, i.e. biosurfactant-mediated transfer and direct interfacial uptake.

Trans-membrane transport of hydrocarbons

Trans-membrane transport of hydrocarbon is the next step following the physical interaction between bacterial cells and hydrocarbons. Before metabolizing the hydrocarbons, microbial cells physically access soluble, emulsified hydrocarbons and big oil droplets, followed by transportation of these substrates across cell membranes (Bouchez-Naitali 2008; Rosenberg 1993) and then create inclusions (Alvarez et al. 1997; De Andres et al. 1991). Further, bacterial biodegradation of hydrocarbons necessitates the passage of hydrophobic substrate across the cell membrane, implying that hydrocarbon trans-membrane transport is the initial step in biodegradation (Hearn et al. 2008). There are three primary ways for transporting substances over the cell membrane as follows: (1) passive diffusion, (2) passive facilitated diffusion, or (3) energy-dependent/active uptake (Hua and Wang 2014; Wang et al. 2020).

Passive diffusion of hydrocarbons

Passive transport refers to the diffusion of matter along the concentration gradient, i.e. from areas of high concentration to the areas of low concentration. Simple diffusion (also known as free diffusion) and alienation diffusion are two types of passive diffusion. In simple diffusion, pollutants are dispersed along a concentration gradient without any support of membrane proteins and need of additional materials during transit (Borbas et al. 2016; Foster and Miklavcic 2016). However, alienation diffusion distinguishes itself by having a high rate of transport that is proportional to the concentration of the material being carried within a given region.

Active transport

Active transport, however, transports the material against a concentration gradient; it is energy requiring process or is associated with a system for releasing energy. It relies on the transport proteins embedded in the membrane and it is selective as well as specific (Vedovato and Gadsby 2014).

Intracellular lipid inclusions formation

Petroleum may usually be integrated into cells via a unique and inducible transport route, which is especially crucial for oil biodegradation. In order to survive in changing environments, microorganisms have evolved to accumulate storage lipids (De Andres 1991; Isken and de Bont 1998). For instance, distinct forms of intracytoplasmic sudanophilic, electron-transparent and electron-dense inclusions were observed in Rhodococcus opacus strain PD630 cells during growth on phenyldecane and gluconate (Alvarez et al. 1996). However, in both the phenyldecane as well as gluconate grown cells, these inclusions primarily had a sphere-shaped form, with disc-shaped inclusions appearing in much more minor numbers. In a like manner, it was reported that when cultivated on n-octadecane, Pseudomonas spp. DG17 has clear-vesicle-type circular inclusions (Hua et al. 2013). Moreover, in different microorganism, the makeup of inclusions, such as lipids and lipophilic chemicals, which represents the inclusion bodies as a carbon and energy storage, differs. For instance, when Rhodococcus opacus strain PD630 cells were grown in phenyldecane as the carbon source, primarily saturated and unsaturated straight long-chain fatty acids (LCFAs) were observed in the inclusion's lipids; however, under any growing conditions, 3-hydroxy alkanoic acids and other hydroxy alkanoic acids were not observed (Alvarez et al. 1996).

Factors affecting bioremediation of petroleum hydrocarbons

The successful implementation of bioremediation techniques requires an understanding of characters and parameters that influence microbial biodegradation of contaminants. Various limiting factors affecting the biodegradation of petroleum hydrocarbons have been identified (Chandra et al. 2013). In addition, bioremediation reactions are regulated by the chemical composition of hydrocarbon and environmental factors such as temperature, nutrients, electron acceptors and substrates which are likely to play significant roles in bioremediation (Varjani and Upasani 2017). Henceforth, numerous studies have discovered that most petroleum hydrocarbon-degrading bacteria yield good results in lab-scale tests but produce poor results in field tests (Head et al. 2006).

Physical and chemical composition of hydrocarbon

While assessing the suitability of a remediation approach, the composition and biodegradability of the petroleum hydrocarbon pollutant must be considered. Excessive petroleum hydrocarbons restrict bacterial growth, resulting in low biodegradation efficiency and even bacterial mortality (Ma et al. 2015). The structure and molecular weight of the hydrocarbon molecule govern its susceptibility to biodegradation. The biodegradation of hydrocarbon is influenced by its bioavailability, which is mostly determined by the concentration, physical state, as well as hydrophobicity, ability to get sorpbed onto soil particles, volatilization and solubility of hydrocarbon. The sensitivity of hydrocarbons to microbial attack varies, and the following order of biodegradability: linear alkanes > branched alkanes low > molecular weight alkyl aromatics > monoaromatics > cyclic alkanes > polyaromatics > asphaltenes (Varjani 2017). This is related to the substrate's physicochemical characteristics and bioavailability, which influence the interaction, transport, as well as transformation of hydrocarbon substrates by bacteria (Varjani and Upasani 2016). However, a wide range of microorganisms degrade aliphatic hydrocarbons and n-alkanes of intermediate chain length (C10–C24). Long-chain alkanes are resistant to biodegradation, while short-chain alkanes (less than C9) are toxic to many microbes but are lost quickly in the atmosphere due to their volatility. Generally, the rate of biodegradation is slowed by branching. Aromatic molecules, particularly polyaromatic hydrocarbons (PAHs), decay slowly, whereas a co-metabolism process can destroy alicyclic chemicals. Although most research at the laboratory scale concentrates on the degradation of a single substrate, petroleum hydrocarbon contaminants in nature are complex. Hence, reproducing laboratory results in real-world applications is challenging. For instance, benzene, toluene and phenol are mineralized efficiently by Pseudomonas putida F1 but in substrate mixtures, toluene and benzene promote phenol biodegradation while phenol inhibits benzene and toluene biodegradation (Abuhamed et al. 2004).

Temperature

Temperature affects the chemistry of pollutants as well as the diversity and physiology of the microbial flora. Hence, it is one of hydrocarbon biodegradation's most significant physical elements. Moreover, the temperature can affect bacterial growth as well as metabolism, the soil matrix and the mode of occurrence of pollutants in the soil, thus indirectly altering biodegradation efficiency (Abed et al. 2015). Further, the solubility of hydrocarbons is influenced by temperature (Foght et al. 1996). Generally, at low temperatures viscosity of the oil increases, but the volatility of the low molecular weight hydrocarbons decreases, resulting in a delay in the commencement of biodegradation (Atlas 1975). The temperature affects the solubility of hydrocarbons; at low temperatures, the solubility of hydrocarbons in water increases (Foght et al. 1996). Despite having a broad temperature profile, biodegradation rates of petroleum hydrocarbon typically drop as temperature decreases, which may be attributed to a reduction in the enzymatic activity (Atlas and Bartha 1972; Gibbs et al. 1975). The surrounding ambient temperature affects the characteristics of spilled oil and microbial activity (Venosa and Zhu 2003). Rates of degradation are generally highest in the temperature range between 15 and 20 °C in marine environments, between 30 and 40 °C in soil environments, while it is between 20 and 30 °C in some freshwater environments (Bartha and Bossert 1984; Cooney 1984).

Moreover, it was confirmed by Li et al. (2017) and Xu et al. (2017) that temperature has a significant impact on biodegradation efficiency and reported that during the biodegradation of C16 alkane, the bacterial strains Acinetobacter spp. JLS1 and Pseudomonas aeruginosa JLC1 isolated from Momoge wetlands in Jilin Province, China, displayed distinct sensitivity to temperature. In a laboratory investigation, Roling et al. (2002a, b, 2004) reported that the petroleum hydrocarbons phenanthrene and dibenzothiophenes were well destroyed. However, similar degradation effects were not observed in a field experiment, which could be due to the temperature range considered in the study. Although significant biodegradation of hydrocarbons has been reported in psychrophilic environments in temperate regions (Pelletier et al. 2004; Delille et al. 2004a, b), the majority of information on hydrocarbon degradation focuses on the activities of mesophiles (Yumoto et al. 2002; Pelletier et al. 2004; Delille et al. 2004a, b). Even though petroleum contamination is acknowledged as a severe hazard to polar habitats, there are a few documented studies on the environmental consequences of oil spills in frigid climates (Delille et al. 2004a, b; Chandra et. al 2013).

Nutrients

Nutrients, particularly nitrogen, phosphorus and iron, play a much more critical role than oxygen in the efficient bioremediation of hydrocarbon pollutants (Cooney 1984). Therefore, when a substantial oil spill occurred in freshwater and marine habitats, there was a sharp rise in the supply of carbon; however, availability of phosphorus and nitrogen became a limiting factor for oil degradation (Atlas 1985). Since a high dose of nutrients can speed up the early pace of oil degradation, thereby cutting down the time it takes to clean up contaminated areas (Oh et al. 2001). Similarly, it has been confirmed by previous studies that nutrient fortification boosts bioremediation by boosting microbial biomass (Sanchez et al. 2000; Margesin and Schinner, 2001; Duncan et al. 2003; Maki et al. 2003; Sarkar et al. 2005). However, on the other hand, excessive nutrient concentrations can sometimes impede biodegradation activity (Chaillan et al. 2006). Likewise, Oudot et al. (1998) and Chaineau et al. (2005) have observed that high NPK levels have a deleterious impact on hydrocarbon biodegradation, particularly on aromatics (Carmichael and Pfaender 1997). Because petroleum hydrocarbons are primarily composed of carbon and hydrogen, the environment must have sufficient additional nutrients such as nitrogen, sulphur, phosphorus, along with various trace elements to support the growth of bacterial degraders (Xu et al. 2018). Converting 1 kg of hydrocarbons in bacterial cells is estimated to reqire approximately 150 g of nitrogen and 30 g of phosphorus (Ron and Rosenberg 2014). On different petroleum-contaminated shorelines and sandy beaches, fertilizers containing bioavailable nitrogen as well as phosphorus have been successfully employed to stimulate petroleum oil biodegradation (Roling et al. 2002a, b; Hazen et al. 2016). In addition to this, several studies have suggested that instead of adding nitrogen supplies to boost bioremediation performance, another good alternative was to use nitrogen-fixing hydrocarbon-degrading bacteria (Thavasi et al. 2006).

Electron acceptors

Hydrocarbons being highly reduced substrates and limited air permeability in petroleum oil-contaminated environments, an electron acceptor in the form of molecular oxygen is crucial for aerobic degradation processes (Xu et al. 2018; Chandra et al. 2013). However, providing ample oxygen to encourage bioremediation of petroleum contaminants in the environment is both costly and impractical. In field studies, Gogoi et al. (2003), found that up to 75% of hydrocarbon pollutants were destroyed within a year by regulating and controlling aeration. Nitrate, iron, bicarbonate, nitrous oxide and sulphate have been reported as alternate electron acceptors during hydrocarbon breakdown in the absence of molecular oxygen. Moreover, encouraging anaerobic microbes by using bulking substances such as sawdust or other electron acceptors (nitrate (NO₃⁻), ferric ion (Fe³⁺) or manganese ion (Mn²⁺)) in anoxic settings are usually more cost-effective than oxygen supplementation (Zedelius et al. 2011; Brown et al. 2017). Even though most studies have shown that hydrocarbon biodegradation is an aerobic process, there have been evidences of anaerobic biodegradation; nonetheless, the former occurs quicker than the latter. The oxidation of the substrate by oxygenases, which requires molecular oxygen, is the first step in the catabolism of aliphatic, aromatic and cyclic hydrocarbons by bacteria and fungi (Singer and Finnerty 1984; Cerniglia 1984; Perry 1984). Therefore, the addition of oxygen can significantly increase the remediation rate. However, the oxygen availability in soil depends upon type of soil, microbial oxygen consumption rates and the presence of utilizable substrates, all of which can lead to oxygen depletion (Bossert and Bartha 1984).

Bioavailability

Bioavailability is the critical factor influencing pollutant biodegradation and refers to the percentage of a pollutant in soil, that is either physio-chemically accessible to microorganisms or altered by microorganisms (Al-Hawash et al. 2018; Chandra et al. 2013). Further, it is termed as the impact of the physical, chemical and microbiological parameters on the rate and extent of biodegradation (Al-Hawash et al. 2018). Since petroleum hydrocarbons are hydrophobic organic pollutants, they possess low bioavailability and low water solubility, making them less prone to photolytic, chemical and biological breakdown (Semple et al. 2003). Biosurfactants and biosurfactant-producing bacteria along with commercially available ionic and non-ionic surfactants possess the ability to enhance bioavailability of organic contaminants (Laha and Luthy 1992; Pennell et al. 1993; Volkering et al. 1993; Miller 1994).

Biosurfactants for bioremediation of petroleum hydrocarbons

Biosurfactants are surface-active, valuable, biologically efficient amphiphilic molecules synthesized by microbes and have a wide range of applications due to their unique characteristics, low toxicity and biological acceptability (Gayathiri et al. 2022, Shivlata et al. 2015). Being an amphiphilic molecule, biosurfactant has monosaccharide, oligosaccharide, polysaccharide and proteins as hydrophilic polar moiety and saturated, unsaturated fatty alcohols or hydroxylated fatty acids as hydrophobic nonpolar moiety (Rodrigues 2015; Santos et al. 2016). However, the hydrophilic–lipophilic balance responsible for determining the hydrophilic and hydrophobic components in surface-active chemicals is one of the essential properties of biosurfactants.

Amphiphilic structure confers to the potential of biosurfactants to change the property of the microbial cell surface and enhance the surface area of hydrophobic substances. Dispersion, emulsification or de-emulsification, wetting, foaming, and coating are some of the features of biosurfactants that make them suitable for physiochemical treatments and bioremediation of various organic as well as metal pollutants (Wu and Lu 2015). Hydrocarbon degraders are widely known for their ability to synthesize biosurfactants in situ, which aid in their survival in environments dominated by hydrophobic compounds (Ganesh and Lin 2009). It is well known that microorganisms use various compounds as carbon and energy sources for their growth and development. Consider an example where the carbon source is an insoluble hydrocarbon. Then, according to Leuchtle et al. (2015), some yeast and bacteria produce various biosurfactants to emulsify hydrocarbons in the medium, although majority of microorganisms change their cell wall structure of their due to the production of lipopolysaccharide in the cell wall (Saenz-Marta et al. 2015). For instance, when the medium contains n-alkanes, Pseudomonas and different species of Torulopsis, respectively, synthesize biosurfactant rhamnolipids and sophorolipids.

Further, Candida lipolytica produces cell-bound lipopolysaccharide, and Acinetobacter species produce emulsan; lipoproteins, like subtilisin, are produced by Bacillus subtilis, while Thiobacillus ferroxidans and Gluconobacter cerinus synthesize ornithinlipids in hydrocarbon-contaminated environment. Karlapudi et al. (2018), stated that Pseudomonas aeruginosa isolated from seawater polluted with oil possesses the ability to break down hexadecane, octadecane, heptadecane and nonadecanes after 28 days of incubation and produces biosurfactants. Besides, Zhuang et al. (2002) reported that range of hydrocarbons like pristine, tetradecane and 2- methylnaphthalene were effectively degraded by Pseudomonas aeruginosa. Moreover, the biosurfactant producing potential of numerous bacterial genera, such as Pseudomonas, Acenetobacter, Bacillus, Rhodococcus, Alcaligenes and Corynebacterium, resulting in efficient petroleum oil degradation, has been widely investigated (Cameotra and Makkar 2004; Abbasian et al. 2016). Degradation of hydrocarbons was studied with inoculation of Acinetobacter haemolyticus and biosurfactant-producing strain Pseudomonas ML2 in soil contaminated with hydrocarbon for an incubation period of 2 months. It was observed that reduction of 39–71% and 11–71% in hydrocarbon concentration was achieved by Acinetobacter haemolyticus and Pseudomonas ML2, respectively (Karlapudi et al. 2018). Another study found that when a chemical surfactant, FinasolOSR-5, was supplemented with a biosurfactant, trehalose-5, dicorynomycolates, the oil degradation capacity of FinasolOSR-5 was multiplied, resulting in the complete removal of aromatic hydrocarbons from contaminated soil within a given period (Itrich et al. 2015). These findings revealed that cell-free biosurfactant of bacterial origin possess remarkable hydrocarbon-degrading potential (Karlapudi et al. 2018). Henceforth, the potential of microbes to bio-remediate hydrocarbons coupled with ability to synthesize biosurfactant can be exploited to speed up the bioremediation of hydrocarbon-contaminated environment (Kumar et al. 2006).

Classification of biosurfactant

The majority of biosurfactants are anionic or neutral; however, those with amine groups are cationic. Biosurfactants have a molar mass ranging from 500 to 1500 Dalton (Da) (Bognolo 1999). However, on the basis of their molecular mass, biosurfactants are classified as low molecular mass biosurfactant and high molecular weight biosurfactant (Saenz-Marta et al. 2015; Harvey 2017). Low molecular weight biosurfactant (i.e. Glycolipids, lipopeptides and phospholipids) reduces surface and interfacial tensions and high molecular weight biosurfactant (i.e. lipoproteins, lipopolysaccharides, amphipathic polysaccharides, polymeric surfactants and particlulate surfactants) are efficient stabilizing agents (Rosas-Galvan et al. 2018; Karlapudi et al. 2018). Biosurfactants are categorised primarily by their source of origin and chemical structure of their hydrophobic components as (1) glycolipid type, (2) fatty acid type, (3) lipopeptide type and (4) polymer type (Gayathiri et al. 2022; Desai and Banat 1997).

Glycolipid

Glycolipids are also known as hydroxy aliphatic acids and are made of long-chain aliphatic acids in association with carbohydrates. An ether or ester group is used to connect the molecules. (Mnif et al. 2018; Sanjana et al. 2017). Rhamnolipids, sophrolipds, trehalolipids and fructose-lipids are examples of glycolipids (Rikalovic et al. 2015; Gayathiri et al. 2022). The rhamnolipid is composed of one or two molecules of rhamnose coupled to one or two molecules of b-hydroxy decanoic acid (Abalos et al. 2001). They are synthesized by Pseudomonas aeruginosa and are a blend of α-L-rhamnopyranosyl-α-L-rhamnopyranosyl-β-hydroxydecanoate (Rha-Rha-C10) and α-L-rhamnopyranosyl-α-L-rhamnopyranosyl-β-hydroxydecanoyl-β-hydroxydecanoate (Rha-Rha-C10-C10) along with their mono-rhamnolipid congeners (Rha-C10-C10 and Rha-C10) (Abdel-Mawgoud et al. 2010). Sophorolipids, on the other hand, are synthesized by yeast of genus Candida (Cortes-Sanchez et al. 2013; Daverey and Pakshirajan 2009] and are made up of a disaccharide sophorose unit linked to long-chain hydroxylated fatty acid by glycosidic bond (Morita et al. 2016). However, trehalolipids are produced by species of Mycobacterium spp., Nocardia spp. and Corynebacterium spp. (Vijayakumar and Saravanan 2015). Trehalose is a disaccharide sugar that is connected to long-chain fatty acids in mycolic acid at the 6th position of the carbon backbone (Karlapudi et al. 2018).

Lipoproteins and lipopeptides

Cyclic lipopeptides consists of a lipid components and 8–17 amino acids with considerable variation in their lipid and amino acid composition (Desai and Banat 1997). Efficient crude oil removal was observed using lipopeptide synthesized by Bacillus atrophaeus (Zeraik and Nitschke 2010). Cyclic acidic lipopeptide, surfactin generated by Bacillus subtilis is the most effective of all known biosurfactants (Arima et al. 1968; Mulligan and Gibbs 2004). Surfactin is composed of a seven-ring amino acid structure that is linked via lactone bond to the fatty acid chain. It has been observed that surfactin lowers the surface tension below 28 milli Newton per meter (mN m⁻¹) (Nguyen and Gotz 2016).

Fatty acids, phospholipids and neutral lipids

Various bacteria and yeast produce fatty acid and phospholipid during growth on n-alkane and are in great demand due to their diverse and useful qualities as biosurfactants (Rosas-Galvan et al. 2018). Acinetobacter spp. creates phosphotidyl ethanolamine-rich vesicles that form optically transparent alkane micro emulsions in water (Santos et al. 2016). Rhodococcus erythropolis was reported to produce phosphatidylethanolamine which has shown a reduced interfacial tension of less than 1 mN m⁻¹ against hexadecane and a criticle micelle concentration of 30 mg/L (Kretschmer 1982; Ansari et al 2009; Rosas-Galvan et al. 2018).

Polymeric biosurfactants

Polymeric biosurfactants are exocellular compounds derived from bacterial species of various genera and comprises a complex mixture of biopolymers such as polysaccharides, lipopolysaccharides and proteins as their primary components (Rosas-Galvan et al. 2018). In polymeric biosurfactants polysaacharides are covalently bonded to fatty acids via o-ester linkages (Desai and Banat 1997; Zokaei et al. 2018). Best polymeric biosurfactants are emulsan, liposan and mannoprotein (Desai and Banat 1997; Lang 2002; Hatha et al. 2007). Moreover, liposan and alasan are among two of the most widely used polysaccharide protein complexes. Most Acinetobacter species generate heteropolysaccharide biosurfactants, which have extracellular polyanionic activity. Emulsan is one of the most effective emulsifying agent as it emulsifies hydrocarbons in water even at a concentration below 0.01 percent. Liposan is a water-soluble extracellular polymeric emulsifier produced by Candida lipolytica, with less than 20% protein content and carbohydrate content more than 80% (Wilton et al. 2016). Mannoproteins account for the major component of the cell wall of Saccharomyces cerevisiae. It is an effective bioemulsifier that can generate stable emulsions with a wide range of hydrocarbons and other chemicals, implying that it might be used as efficient cleaning agents (Cameron et al. 1988; Gayathiri et al. 2022). The structural mannoproteins, composed of mannopyranosyl linked to a tiny protein, are the most common, while the enzymatic mannoproteins, which have more protein moieties, are the most effective emulsifiers. They stimulate the synthesis of antibodies by immunological cells (Casanova et al. 1992; Oliveira et al. 2009b).

Particulate biosurfactant

Particulate biosurfactants generate a microemulsion by partitioning extracellular vesicles and fimbriae, which influences alkane absorption in microbial cells (Gayathiri et al. 2022; Santos et al. 2016). For example, the vesicles synthesized by Acinetobacter spp are proteins, phospholipids and lipopolysaccharides with a diameter and buoyant density of 20 to 50 nanometer (nm) and of 1.158 cubic gram per centimetre (g³cm⁻¹), respectively (Vijayakumar and Saravanan 2015; Chakrabarti 2012). Different Biosurfactants produced by respective microorganisms and their surface tensions values are listed in Table 2.

Table 2.

Surface tension values & biosurfactants produced by selected bacterial strains

Biosurfactant	Microorganism	Surface tension (mN m−1)	Property	References
Rhamnolipids	Pseudomonas aeruginosa L2-1, Bacillus spp. AB-2	29	Playing a crucial role in the field of pharmaceuticals	Magalhes and Nitschke (2013) and Amani et al. (2013)
Sophorolipid	Candida lipolytica Y-917, Torulopsis bombicola	33	Produces hydrocarbon and oil emulsions in a liquid like water	Alizadeh-Sani et al. (2018); Imura et al. (2014)
Mannan-fatty acid	Candida tropicalis	30	Recognized as key antigenic determinants	Kuraoka et al. (2020) and Chen et al. (2011)
Surfactin	Bacillus subtilis ATCC 21,332	27–32	Used in enhancement of the iron-remediation, anti-inflammatory activity	Wang et al. (2005), Yea et al. (2019) and Liu et al. (2015)
Trehalose, sucrose, and fructose, lipids	Arthrobacter spp., Rhodococcus aurantiacus	36	Lower the interfacial tension and make hydrophobic compounds more “pseudosoluble.”	Franzetti et al. (2010) and Kuyukina et al. (2015)
Lipopeptide	Arthrobacter MIS 38 Pseudomonas fluorescence, Bacillus atrophaeus	27	Low interphase surface tension due to emulsifying action	Zhang et al. (2015), Yakimov et al. (1995)
Liposan	Candida lipolytica	29	Produces stable oil/water emulsions	Campos et al. (2013)

Adapted from: Karlapudi et al. (2018) and Gayathiri et al. (2022)

Evidences of hydrocarbon degradation by biosurfactant synthesizing bacteria

Hydrocarbons are difficult to degrade because of their hydrophobic properties, which enable them to adhere to a variety of substrates. Being amphiphillic, biosurfactants possess both hydrophilic and hydrophobic moieties, which imparts them surface-active properties that lower surface and interfacial tension in aqueous solutions and mixtures of hydrocarbons (De Almeida 2016; Santos et al. 2017, 2016). This suggests that biosurfactants and microorganisms that produce biosurfactants can aid in the detoxification of petroleum hydrocarbon from contaminated environments (Deng et al. 2020; Jadeja et al. 2019; Joe et al. 2019; Ravindran et al. 2020). Biosurfactants are microbial compounds with a high surface activity and emulsification ability which can sustain high temperatures and salt concentrations, enabling them to remain stable in extreme environments (Bami et al. 2022). Pseudomonas and Bacillus are, however, the most often reported biosurfactant producing bacterial strains (Ayed et al. 2015). Serratia marcescens ZCF25 and Bacillus cereus UCP 1615, both isolated from oily sludge synthesize highly stable lipopeptide type biosurfactants which reduce surface tension and possess the potential to clean up oil spills (Durval et al. 2020; Huang et al. 2020). In addition to this, Patel and Patel (2020), confirmed that biosurfactants produced by Stenotrophomonas spp. S1VKR-26 can be employed for remediation of petroleum contamination in marine ecosystems. Biosurfactants derived from Rhodococcus erythropolis HX-2 have been found to improve petroleum biodegradation by improving the hydrophilic compound's solubility (Hu et al. 2020).

According to Ambust et al. (2021), the biosurfactant generated by Pseudomonas spp. SA3 has emulsification and surface tension reduction ability of 43% and 34.5 (milli Newton per meter) mN m⁻¹, respectively, encouraging growth of agricultural crop in petroleum-contaminated soils. It has been reported that Serratia spp., obtained from a petroleum-contaminated site, when cultured in the presence of used vegetable oil, creates a biosurfactant that enhances the solubility of variety of contaminants such as tetrachloroethylene (TCE), perchloroethylene (PCE), naphthalene, toluene and phenanthrene (Liu et al. 2009; Vipulanandan and Ren 2000; Harendra and Vipulanandan 2011, 2008; Kim and Vipulanandan 2005). Aqib Hassan et al. (2017) compared the hydrocarbon degradation potential of four biosurfactant-producing bacterial strains isolated from petroleum hydrocarbon-contaminated soil (Pseudomonas poae BA1, Acinetobacter bouvetii BP18, Bacillus thuringiensis BG3 and Stenotrophomonas rhizophila BG32) with non biosurfactant-producing hydrocarbon degrading Pseudomonas rhizosphaerae strain BP3. They confirmed that biosurfactant producing strains displayed a 16–28% increase in hydrocarbon breakdown when compared to non-biosurfactant producing bacteria. Furthermore, they stated that as compared to non biosurfactant-producing strain biosurfactant-producing strains showed 16–28% greater hydrocarbon degradation potential, as compared to non biosurfactant-producing strain. Additionally, BA1 showed the highest hydrocarbon degradation potential, i.e. 96.07%, followed by BP18 (93.53%), BG3 (89.97%), BG32 (87.10%) and BP3 (74.60%). It can be inferred that factors such as surface tension, hydrophobicity of cell and biosurfactant synthesis affect degradation of hydrocarbons. Further, reduction in hydrophobicity of hydrocarbons and subsequent increase in biodegradation can be achieved by employing microbial biosurfactants. Table 3 reveals several studies on biosurfactant-assisted remediation of petroleum hydrocarbons.

Table 3.

Evidences of biosurfactants enhanced remediation of petroleum hydrocarbons

Microorganisms	Pollutant type	Biosurfactant type	Mode of action	References
Achromobacter spp. A-8	Petroleum	Biosurfactant (not specified)	High performance in displacement of oil by minimizing surface tension	Deng et al. (2020)
Acinetobacter sp. D3-2	Petroleum	Lipopeptide	Reduces surface tension from 48.02 to 26.30 mN m−1, thus degrading 82% of hydrocarbons	Bao et al. (2014)
Serratia spp.	Hydrocarbon	Lipopeptide	Increases surface area of hydrocarbons by reducing surface and interfacial tension, thus making them easily accessible to the bacteria	Gidudu et al. (2020)
Bacillus amyloliquefaciens An6	Diesel oil	Biosurfactant An6	Reduce surface tension from 72 mN m−1 to below 30 mN m−1 with a CMC of 100 mg/L	Ayed et al. (2015)
Wickerhamomyces anomalous	Crude oil	Lipopeptide	Surface tension reduction	Souza et al. (2018)
Pseudomonas aeruginosa AT10	Total petroleum hydrocarbons, the group of isoprenoids from the aliphatic fraction and the alkylated PAHs from the aromatic fraction	Rhamnolipid MAT10	Surface tension & interfacial tension is reduced up to 26.8 mN m−1 & 1 mN m−1 respectively, with CMC of 150 mg/L	Geddes et al. (2008)
Bacillus algicola (003-Phe1), Rhodo- coccus soli (102-Na5), Isoptericola chiayiensis (103-Na4), and Pseu- doalteromonas agarivorans (SDRB-Py1)	Crude oil	Rhamnolipid	Maximizing crude oil emulsification	Lee et al. (2018)
Serratia marcescens UCP 1549	Burned motor oil	Lipopeptide	Decreases surface tension & emulsifies oil, improves solubility of oil in water	Araujo et al. (2019)
Mixed consortia of Pseudomonas aeruginosa strain PG201, Rhodococcus spp. H131A, and a Pseudomonas strain	Hexadecane, dodecane, benzene, toluene, iso-octane, pristane (2,6,10,14-tetramethyl pentadecane), naphthalene, and phenanthrene	Rhamnolipid Dyna270	Biosurfactant enables linear alkane to degrade at faster rate than monoaromatics	Inakollu et al. (2004)
Paenibacillus spp. D9	Diesel and motor oil	Biosurfactant (not specified)	Reduction in surface tension up to 31.2 MN m−1 increases solubility of contaminants causing 74.3% &77.6% removal of diesel & motor oil respectively	Jimoh and Lin (2020)
Bacillus stratospheric strain FLU	Motor oil	Lipopeptide	Increases solubility of motor oil by promoting micelles formation	Nogueira Felix et al. (2019)

Adapted from: Wang et al. (2020) and Bami et al. (2022)

CMC critical micelle concentration, TPH total petroleum hydrocarbon, PAH polycyclic aromatic hydrocarbon, mN m-1 milli Newton per meter

Mode of action of biosurfactants in bioremediation of petroleum hydrocarbon

The bioavailability of contaminants to microorganisms is an essential factor influencing the microbial degradation process (Lawniczak et al. 2013). Microorganisms produce biosurfactants through excretion or attachment to cells, especially when cultured on water-insoluble substrates (Santos et al. 2016). By improving the solubility of hydrophobic organic contaminants, biosurfactants improve their bioavailability to microorganism for biodegradation (Kreling et al. 2020). At bulk concentrations over the critical micelle concentration (CMC), biosurfactants with an amphiphilic structure aggregate and form micelles in the hydrophilic environment because a thermodynamically stable micelle structure forms at equilibrium. In micelle formations, hydrophobic groups of surfactants orient toward the hydrophobic environment, while hydrophilic groups orient toward the aqueous phase; as a result, hydrophobic contamination disperses and dissolves in the aqueous solvent. Micelle formation, therefore, speeds up the absorption of substances by microbial cells (Effendi et al. 2018; Jahan et al. 2020; Nogueira Felix et al. 2019). Degradation of hydrocarbon is aided by biosurfactants in the following two ways: (i) making organic hydrophobic contaminants more accessible to microbial attack by emulsifying them and enhancing their solubility; (ii) decreasing the surface tension between bacterial cells and petroleum compounds via regulating the interaction between the hydrophobic substrate and microbial cell surface, thereby allowing the bacterial cells to associate easily with the hydrophobic contaminants (Effendi et al. 2018; Pacwa-Płociniczak et al. 2011). Figure 2 depicts the mechanisms of biosurfactant-assisted microbial breakdown of hydrophobic compounds. Further, the adhesion of microbes to hydrophobic surfaces is largely determined by their hydrophobicity of bacterial cell surfaces, which is a growth limiting factor in environments containing hydrophobic substrates (Zhang and Miller 1994; Zhao et al. 2011). Microorganisms capable of lowering surface tension to less than 40 dynes per centimetre (dyne cm⁻¹) are considered as efficient biosurfactant producers (Ruggeri et al. 2009). An effective biosurfactant, according to Seydlova and Svobodova (2008), minimizes surface tension of water significantly from 72 to 27 dyne cm⁻¹. According to Ben Ayed et al. (2014), biosurfactant produced by Bacillus mojavensis A21 improves the solubility of hydrophobic organic compounds by lowering surface resistance and increasing surface activity at the air/water interface, making it easier to partition hydrocarbon from the hydrophobic oil phase into the aqueous solution. Concentration of hydrocarbon pollutant further influences solubilization efficiency, and a high hydrocarbon content might thus be a limiting factor in the solubilization process.

Fig. 2.

Mechanisms of biosurfactant-assisted microbial breakdown of hydrophobic compounds. Adapted and redrawn from: Bami et al. (2022)

Pathways for biosurfactant production

Depending on the principal carbon source used in the culture medium, a precursor for Biosurfactant is synthesized via different metabolic pathways. For example, when carbohydrates constitute the only carbon source for the manufacture of a glycolipid, the carbon flow is managed in such a manner that the microbial metabolism suppresses both hydrophilic moiety formation (glycolytic pathway) as well as lipid formation (lipogenic pathway) (Haritash and Kaushik 2009). Degradation of the hydrophilic substrate, such as glucose or glycerol, occurs until intermediates of the glycolytic pathway, such as glucose 6-phosphate are formed, which is one of the main precursors of carbohydrates present in a hydrophilic moiety of biosurfactant. However, for lipid production, glucose is first oxidized to pyruvate by glycolysis, followed by the transformation of pyruvate to acetyl-CoA, which generates malonyl-CoA when combined with oxaloacetate. Finally, malonyl-CoA is converted into a fatty acid, the substrate for synthesizing the lipids (Hommel and Huse 1993; Santos et al. 2016). However, when the carbon source is a hydrocarbon, lipolytic pathway and gluconeogenesis (the creation of glucose from various hexose precursors) occur, allowing hydrocarbon compounds to be used for the fatty acids or sugar production, respectively. Microorganisms use hydrophilic substrates for cell metabolism and the synthesis of the polar moiety of a biosurfactant, whereas hydrophobic substrates are used mainly for the manufacture of the biosurfactant’s hydrocarbon component (Desai and Banat 1997; Weber et al. 1992).

For cell metabolism and synthesis of a polar moiety of a biosurfactant, microorganisms primarily use hydrophilic substrate; however, for the synthesis of a nonpolar moiety of biosurfactant, the hydrophobic substrate is utilized exclusively. Biosynthesis of a surfactant occurs through the following four different routes: (Sydatk and Wagner 1987):

1. Carbohydrate and lipid synthesis.

2. Synthesis of the carbohydrate half while the synthesis of the lipid half depends on the length of the chain of the carbon substrate in the medium.

3. Synthesis of the lipid half while the synthesis of the carbon half depends on the substrate employed.

4. Synthesis of the carbon and lipid halves, both dependent on the substrate.

As a result, the length of the n-alkane chain employed as a carbon source influences surfactant biosynthesis. For example, the formation of manosilerythritol lipid (MEL) by the yeast Candida antarctica in various n-alkanes was discovered by Kitamoto et al. (2001) and discovered that this species does not grow or create a biosurfactant in C10 to C18 media. However, the highest yield was obtained when the species was cultivated in a medium containing C12 to C18, with octadecane as the substrate. Figure 3 depicts the schematic diagram of the biosynthetic pathway for biosurfactant production.

Fig. 3.

Schematic diagram of biosynthetic pathway for synthesis of biosurfactant. Adapted and redrawn from: Karlapudi et al. (2018)

Biosurfactant’s influence on bioavailability of organic hydrocarbon contaminants

The following mechanisms can help biosurfactants to improve the bioavailability of hydrophobic hydrocarbons: emulsification of liquid pollutants in the non-aqueous phase (Volkering et al. 1998; Bustamante et al. 2012); augmentation of the contaminant’s solubility (Volkering et al. 1995; Edwards et al. 1991; Bustamante et al. 2012) and facilitating pollutant transfer from solid matrix (Bustamante et al. 2012; Yeom et al. 1996). However, the physical state of pollutants determines the proportional contribution of these mechanisms to increased mass transportation (Volkering et al. 1998). In addition to above, Bustamante et al. (2012) mentioned a fourth probable mechanism, i.e. biosurfactants aid microorganisms in adsorbing contaminant-occupied soil particles, thereby reducing the diffusion route length between the sites of adsorption and bio-uptake by the microorganisms. Biosurfactants can help in bioremediation by mobilizing, solubilizing, or emulsifying contaminants (Urum and Pekdemir 2004; Nguyen et al. 2008). Biosurfactants with a low molar mass stimulate mobilisation and solubilization mechanisms below and above the CMC, respectively. However, the emulsification process is aided by a biosurfactant with a high molecular mass (Urum and Pekdemir 2004; Pacwa-Plociniczak et al. 2011).

Biosurfactant-mediated emulsification and solubilization of hydrophobic organic contaminants

This mechanism involves biosurfactant aided lowering of interfacial tension between aqueous and non-aqueous phase, hence leading to the development of micro and macro emulsions. As a result, the contact area increases, allowing for enhanced mass transfer of pollutants to the aqueous phase as well as mobilization of sorbed liquid-phase contaminants (Volkering et al. 1998; Edwards et al. 1991).

Augmentation of the contaminant’s solubility

The presence of micelles with large quantities of organic molecules in the hydrophobic centre of the micelles increases the apparent solubility of hydrophobic organic contaminants (Volkering et al. 1995; Edwards et al. 1991). According to Brown (2007), contaminant's apparent aqueous solubility is defined as the sum of aqueous phase hydrocarbon concentrations (Caq) and micellar phase hydrocarbon concentrations (Cmic).

Enhanced transportation of the contaminants from the solid phase

According to Volkering et al. (1998), this mechanism encompasses a number of strategies, including the pollutant’s interaction with individual biosurfactant moiety, the interaction between surfactants and separate-phase or sorbed hydrocarbons, the mobilisation of contaminants by organic matrix swelling and the mobilisation of contaminants trapped in soil by a decreasing surface tension of the water in the pores of soil particles.

Nanoparticles: a promising tool for enhanced biosurfactant production

The metabolic constraints of the pathways involved in biosurfactant synthesis limit biosurfactant production in bacteria (Desai and Banat 1997). According to Kiran et al. (2011) metals are one of the crucial components that influence the biosurfactant synthesis by microorganisms. Among the several metal ions, iron (Fe) is usually regarded as a critical nutrient required by different microbes for the formation of biosurfactants (Haferburg and Kothe 2007). Due to its natural bio-compatibility (Perez et al. 2002) and low toxicity (Liu et al. 2013), the Fe atom is commonly utilised by microbes in various key metabolic pathways. For instance, Kiran et al. (2014) observed that in comparison to control (absence of Fe nanoparticles), Nocardiopsis MSA13A produces 80% more biosurfactant in the presence of Fe-nanoparticle (10 milligrammes per litre; mg L⁻¹). Furthermore, no effect of Fe on the morphology of the filamentous structure of Nocardiopsis MSA13A was discovered using a scanning electron microscope, indicating that Fe nanoparticles are non-toxic to the bacteria. In a similar study conducted by Liu et al. (2013) in Serratia spp. approximately 63% rise in the biosurfactant production and 57% increase in bacterial growth were observed in the presence of just 1.0 mg L⁻¹ of Fe nanoparticle. However, larger concentrations of Fe-NP impede the growth of the synthesizing microorganism, indicating the importance of restricting Fe ions in the culture media for adequate production of glycolipid biosurfactant (Liu et al. 2013). Later, it was found that adding 1 mg L⁻¹ of iron-silica nanoparticles (Fe-Si-NP) to a Pseudomonas aeruginosa strain for 6 h increases production of rhamnolipid by 57% as compared to a medium without nanoparticles (Sahebnazar et al. 2018). This rise in yield of biosurfactant was ascribed to increased cell proliferation and subsequent cell lysis at higher NP concentrations, which released the biosurfactant into the culture medium. Hence, considering above studies, it can be inferred that employing low concentrations of nanoparticle, such as iron nanoparticles, silica nanoparticles can be an emerging strategy for improving biosurfactant synthesis.

Nanoparticle-mediated bioremediation of petroleum hydrocarbons

Lower solubility of the hydrophobic elements of crude oil and their bioavailability to microorganisms hinders the natural process of petroleum hydrocarbons bioremediation. At the same time, the hazardous and hydrophobic property of hydrocarbons in crude oil leads to poor microbial biomass sustainability, especially at increasing petroleum hydrocarbon concentrations. Hence, in order to overcome its perceived limitations, the current state-of-the-art in the field of nanotechnology offers multiple ways for its possible applicability in bioremediation technology. Nanomaterials are made up of particles with at least one dimension between 1.0 and 100 nm (nm). Some particular attributes of nanoparticles, such as high surface-to-volume ratio, improved magnetic properties and specialized catalytic characteristics impart considerable significance to these nanoparticles (NPs) than their bulk phase counterparts for remediating the pollutants at enhanced rate with generation of least amount of toxic secondary residues (Gupta et al. 2011; Bhattacharya et al. 2013). According to Mehndiratta et al. (2013), remediation of potentially harmful pollutants can be accomplished with a variety of nanomaterials such as nanoscale zeolites, multi-walled carbon nanotubes, biometallic particles (iron nanoparticle), dendrimer enzymes and metal oxides. Ansari and Husain (2012) stated that, owing to the fact that NPs create a neutral and biocompatible microenvironment that do not interfere with the enzyme’s inherent properties and assist in the preservation of their biological functions, the use of nanoparticles in enzyme-mediated remediation technology is getting popular. Furthermore, the magnetic properties of NPs makes it easy to separate immobilised enzymes or proteins from reaction mixtures by generating magnetic field. As a result, centrifugation or filtration are not needed, which would otherwise prolong the procedure and operational challenges (Ranjbakhsh et al. 2012; Khoshnevisan et al. 2011). Moreover, nanoparticles assist in adsorption and chemical oxidation of petroleum contaminants, thus indirectly augmenting the bioremediation process.

Adsorption of petroleum hydrocarbon contaminants by NPs

Nanoparticle adsorption capability decreases the net harmful impacts of soil and water pollutants and facilitates growth of microorganisms even at greater toxicant concentrations (Nowack and Bucheli 2007). It has been stated by Masooleh et al. (2010) that hydrocarbons such as gasoline, kerosene, toluene and crude oil can be adsorbed by organo-clay upto five times of its mass. Surfactants such as quaternary ammonium cations, which come in the form of [(CH₃)₃NR]⁺ or [(CH₃)₂NRR']⁺ (where R' or R denotes an aromatic or alkyl group), replace inorganic cations and increase the hydrophobicity of clay, leading to the creation of organophilic clay (Safaei et al. 2008). Gotovac et al. (2006) and Yang et al. (2006) suggest thar Carbon nanotubes (CNT) have been employed as an effective polycyclic aromatic hydrocarbon (PAH) adsorbent. However, other organic pollutants such as chlorophenol, dichloro diphenyl trichloroethane (DDT), phthalateesters can be adsorbed using CNT (Nowack and Bucheli 2007). Hybrid nanoparticles such as Magnetic Shell Cross-linked Knedel-like nanoparticles (MSCKs), created by co-assembling hydrophilic Poly Styrene (PAA20-b-PS20) with hydrophobic Poly-Acrylic Acid block copolymers and tetrahydrofuran with oleic acid stabilised Fe₂O₃ nanoparticles are efficient hydrocarbon sequestering agents from crude oil (Pavia-Sander et al. 2013). Further it was discovered that these nanoparticles eliminate hydrophobic hydrocarbon pollution approximately ten times of their weight (Pavia-Sanders et al. 2013). In a similar study, Hu et al. (2014) exhibited oil sorption using compressible carbon nanotube–graphene hybrid aerogels possessing ultra oleophilicity and hydrophobicity. In addition to this, a hydrophobic and non-biodegradable constituent found in crude oil, i.e. asphaltene, has been reported to get adsorbed using ferric ocide (Fe₂O₃₎ and zinc oxide (ZnO) nanoparticles. Over the past few years, various types of nanoparticles have been successfully explored for remediation procedures. For instance, owing to its high reactivity, low toxicity, low cost, along with wider range of applications, nanoscale zero valent iron (nZVI) is one of the most extensively utilised nanoparticles and hence can be successfully employed for eradicating polycyclic aromatic hydrocarbons (PAHs) and total petroleum hydrocarbons (TPHs) from contaminated soils (Vu and Mulligan 2020). Therefore, combining nanoparticles and biosurfactants will be a potential strategy for remediation of petroleum-contaminated soils.

NP-mediated chemical oxidation of petroleum hydrocarbons

Nanoparticles can promote hydrocarbon bioremediation by facilitating oxidation of these chemicals, minimizing their detrimental effects and stimulating microbial growth. Nano-peroxide or calcium peroxide (CaO₂) nanoparticles are effective oxidants and regarded as a reliable oxygen releasing medium, thus accelerating the solubility and biodegradation rate of hydrocarbons (Northup and Cassidy 2008). In this regard, Pereira and Hanna (2005) reported that calcium peroxide (CaO₂) nanoparticles can substantially oxidize mixture of gasoline and benzene (up to 800 mg L⁻¹) within 24 h. Further, hydrogen peroxide (H₂O₂) and iron oxide (FeO) when mixed in 33.7:1 ratio, eliminate 91% of the total petroleum hydrocarbon in the time span of 4 h (Ershadi et al. 2011). Moreover, to remove organic contaminants from petroleum refinery effluent, Saien and Shahrezaei (2012) employed nano-titanium oxide as a photocatalyst in the presence of UV light. It was observed that 60 to 90 min’ exposure to UV-rays, at a temperature of 45 °C and pH 3.0, a very small concentration of nano-titanium oxide catalyst, i.e. 100 mg L⁻¹ could destroy roughly 78% of the organic pollutant. The abovementioned study was further supported by Fard et al. (2013) and Ziollia and Jardium (2001), who, respectively, employed film of titanium dioxide (TiO₂₎ nano-powder layer for elimimating petroleum hydrocarbons and colloidal TiO₂ nanoparticles for photo-catalytic disintegration of soluble crude oil portion in contaminated seawater.

Nanoparticles: issues and future prospects

In the domain of bioremediation of petroleum hydrocarbon-polluted environments, nanotechnology provides additional benefits. However, the unintended or accidental release of nanoparticles into the environment has a negative impact on ecosystem sustainability, underlining the urgent need to eliminate nanoparticles from the environment (Handy et al. 2008). Moreover, nanomaterials, if released into the environment uncontrollably, may be dangerous to microorganisms involved in the clean-up process and hence it is vital to select nanomaterial appropriately. For instance, nanomaterials such as AgO, Fe₂O₃, CuO, ZnO, etc. are noted for their antimicrobial as well as toxicological properties, when present in excess (Azam et al. 2012; Zhou et al. 2012). In this context it has been observed by Stampoulis et al. (2009) that silver (Ag) nanoparticles cause plant biomass reduction and copper (Cu) nanoparticles cause root length reduction, due to their poisonous characteristics. Moreover, due to the lack of cost effectiveness and eco-friendliness of physical and chemical procedures for eradicating these nanoparticles from soil and water, nanomaterial may not be practicable. Hence, it is a must to analyse the ecological risk and repercussions generated by nanoparticles in order to acquire complete understanding of the environmental fate of these nanoparticle and to design a cost-effective, environmentally sustainable and target-oriented bioremediation strategy.

Conclusion

Waste from the petroleum industry contains various hydrocarbons that are harmful to human health, biological diversity and the environment as a whole. Bioremediation employing potential hydrocarbon-degrading microorganisms overcomes the well-known constraints of conventional physical and chemical treatment approaches. On the other hand, Biosurfactants may be an appealing choice for improving bioremediation efficiency due to their versatility, biodegradability, ecological safety and environmental acceptance. In order to break down organic contaminants like petroleum, biosurfactants can be utilized as a low-cost solution that does not require specialized in-situ techniques or other equipment. By emulsifying hydrophobic pollutants, biosurfactants boost hydrocarbon pollutant’s solubility and allow microorganisms more access to contaminants through the formation of the micelle, resulting in contamination clearance without forming a new hazardous metabolite. Although the biosurfactant sector has grown dramatically in recent decades, large-scale synthesis of these biomolecules remains a problem due to the vast disparity between the required financial investment and commercially viable industrial production. Further, nanotechnology may be a potential approach for overcoming the limitations of the bioremediation process and ensuring large-scale production of biosurfactants. However, since these nanoparticles might affect the living biota, it is critical to investigate their toxicological and environmental effects prior to their application at field scale. Henceforth, still there is a need to conduct a number of experiments before biosurfactant- and nanoparticle-assisted bioremediation may be successfully applied for the reclamation of petroleum-contaminated settings.

Future prospects

From this review, it can be inferred that microbial degradation of petroleum hydrocarbons could be an efficient clean-up strategy for petroleum hydrocarbon contamination. Better understanding of petroleum-degrading microbial communities and the underlying petroleum biodegradation mechanisms could, however, assist in predicting the fate of petroleum hydrocarbon compounds in the ecosystem, as well as in designing viable petroleum hydrocarbon bioremediation techniques. Furthermore, owing to the limited knowledge of various mechanisms involved in bioremediation along with the regulation of enzyme system degrading pollutants, maximum potential of wild species could not be realised, so the use of genetically modified bacteria (GM bacteria) represents a potential scientific area with far-reaching implications. Although biosurfactant-assisted bioremediation of petroleum hydrocarbon-contaminated sites appears to be an appealing solution, however, biosurfactant's expensive manufacturing and purifying costs, as well as poor yield, limit its use in this field. Additionally, biosurfactants are not yet cost effective as compared to chemical surfactants; hence in order to lower the production cost, extensive research into large-scale production of biosurfactants from low-cost renewable substrates, innovative purification techniques, genetic and metabolic engineering tools is needed. Despite indications that employing nanoparticles can be an environmentally acceptable approach for overcoming the limitation of remediation process as well as for enhancing biosurfactant production, minimal research has been done in this domain so far. Hence, in order to ensure that nano-enhanced bioremediation is an environmentally and economically viable strategy to remediate petroleum-contaminated sites and to speed up biosurfactant production, future research should concentrate on larger-scale studies, such as the environmental and biological effects of introducing nanoparticles into the ecosystem.

Authors contribution

Methodology, conceptualization, validation and investigation: DS, AG & MC; Original draft preparation & writing: DS & AG; Review and editing: JPNR. All authors have read and agreed to publish this version of the manuscript.

Funding

Senior Research Fellowship to Diksha Sah from Indian Council of Agricultural Research (ICAR), New Delhi, Goverment of India.

Declarations

Conflict of interest

The authors declare no conflict of interest.