The article is part of the research topic “Advanced bioremediation technologies and synthetic organic compounds (SOC) recycling processes”. View all 14 articles
       Low molecular weight polycyclic aromatic hydrocarbons (PAHs) such as naphthalene and substituted naphthalenes (methylnaphthalene, naphthoic acid, 1-naphthyl-N-methylcarbamate, etc.) are widely used in various industries and are genotoxic, mutagenic and/or carcinogenic to organisms. These synthetic organic compounds (SOCs) or xenobiotics are considered priority pollutants and pose a serious threat to the global environment and public health. The intensity of human activities (e.g. coal gasification, oil refining, vehicle emissions and agricultural applications) determines the concentration, fate and transport of these ubiquitous and persistent compounds. In addition to physical and chemical treatment/removal methods, green and environmentally friendly technologies such as bioremediation, which utilize microorganisms capable of completely degrading POCs or converting them into non-toxic by-products, have emerged as a safe, cost-effective and promising alternative. Various bacterial species belonging to the phyla Proteobacteria (Pseudomonas, Pseudomonas, Comamonas, Burkholderia, and Neosphingobacterium), Firmicutes (Bacillus and Paenibacillus), and Actinobacteria (Rhodococcus and Arthrobacter) in the soil microbiota have demonstrated the ability to degrade various organic compounds. Metabolic studies, genomics, and metagenomic analysis help us understand the catabolic complexity and diversity present in these simple life forms, which can be further applied for efficient biodegradation. The long-term existence of PAHs has resulted in the emergence of novel degradation phenotypes through horizontal gene transfer using genetic elements such as plasmids, transposons, bacteriophages, genomic islands, and integrative conjugative elements. Systems biology and genetic engineering of specific isolates or model communities (consortia) can enable comprehensive, rapid and efficient bioremediation of these PAHs through synergistic effects. In this review, we focus on the different metabolic pathways and diversity, genetic composition and diversity, and cellular responses/adaptations of naphthalene and substituted naphthalene-degrading bacteria. This will provide ecological information for field application and strain optimization for efficient bioremediation.
       Rapid development of industries (petrochemicals, agriculture, pharmaceuticals, textile dyes, cosmetics, etc.) has contributed to global economic prosperity and improved living standards. This exponential development has resulted in the production of a large number of synthetic organic compounds (SOCs), which are used to manufacture various products. These foreign compounds or SOCs include polycyclic aromatic hydrocarbons (PAHs), pesticides, herbicides, plasticizers, dyes, pharmaceuticals, organophosphates, flame retardants, volatile organic solvents, etc. They are emitted into the atmosphere, aquatic and terrestrial ecosystems where they have multidimensional impacts, causing detrimental effects on various bioforms through alteration of physicochemical properties and community structure (Petrie et al., 2015; Bernhardt et al., 2017; Sarkar et al., 2020). Many aromatic pollutants have strong and destructive impacts on many intact ecosystems/biodiversity hotspots (e.g. coral reefs, Arctic/Antarctic ice sheets, high mountain lakes, deep-sea sediments, etc.) (Jones 2010; Beyer et al. 2020; Nordborg et al. 2020). Recent geomicrobiological studies have shown that the deposition of synthetic organic matter (e.g. aromatic pollutants) and their derivatives on the surfaces of artificial structures (built environment) (e.g. cultural heritage sites and monuments made of granite, stone, wood and metal) accelerates their degradation (Gadd 2017; Liu et al. 2018). Human activities can intensify and worsen the biological degradation of monuments and buildings through air pollution and climate change (Liu et al. 2020). These organic contaminants react with water vapor in the atmosphere and settle on the structure, causing physical and chemical degradation of the material. Biodegradation is widely recognized as undesirable changes in the appearance and properties of materials caused by living organisms that affect their preservation (Pochon and Jaton, 1967). Further microbial action (metabolism) of these compounds can reduce structural integrity, conservation effectiveness and cultural value (Gadd, 2017; Liu et al., 2018). On the other hand, in some cases, microbial adaptation to and response to these structures has been found to be beneficial as they form biofilms and other protective crusts that reduce the rate of decay/decomposition (Martino, 2016). Therefore, the development of effective long-term sustainable conservation strategies for stone, metal and wood monuments requires a thorough understanding of the key processes involved in this process. Compared with natural processes (geological processes, forest fires, volcanic eruptions, plant and bacterial reactions), human activities result in the release of large volumes of polycyclic aromatic hydrocarbons (PAHs) and other organic carbon (OC) into ecosystems. Many PAHs used in agriculture (insecticides and pesticides such as DDT, atrazine, carbaryl, pentachlorophenol, etc.), industry (crude oil, oil sludge/waste, petroleum-derived plastics, PCBs, plasticizers, detergents, disinfectants, fumigants, fragrances and preservatives), personal care products (sunscreens, disinfectants, insect repellents and polycyclic musks) and munitions (explosives such as 2,4,6-TNT) are potential xenobiotics that may impact planetary health (Srogi, 2007; Vamsee-Krishna and Phale, 2008; Petrie et al., 2015). This list can be expanded to include petroleum-derived compounds (fuel oils, lubricants, asphaltenes), high molecular weight bioplastics, and ionic liquids (Amde et al., 2015). Table 1 lists various aromatic pollutants and their applications in various industries. In recent years, anthropogenic emissions of volatile organic compounds, as well as carbon dioxide and other greenhouse gases, have begun to increase (Dvorak et al., 2017). However, anthropogenic impacts significantly exceed natural ones. In addition, we found that a number of SOCs persist in many environmental environments and have been identified as emerging pollutants with adverse effects on biomes (Figure 1). Environmental agencies such as the United States Environmental Protection Agency (USEPA) have included many of these pollutants in their priority list due to their cytotoxic, genotoxic, mutagenic, and carcinogenic properties. Therefore, strict disposal regulations and effective strategies for waste treatment/removal from contaminated ecosystems are needed. Various physical and chemical treatment methods such as pyrolysis, oxidative thermal treatment, air aeration, landfilling, incineration, etc. are ineffective and costly and generate corrosive, toxic and difficult to treat by-products. With increasing global environmental awareness, microorganisms capable of degrading these pollutants and their derivatives (such as halogenated, nitro, alkyl and/or methyl) are attracting increasing attention (Fennell et al., 2004; Haritash and Kaushik, 2009; Phale et al., 2020; Sarkar et al., 2020; Schwanemann et al., 2020). The use of these indigenous candidate microorganisms alone or in mixed cultures (colonies) for the removal of aromatic pollutants has advantages in terms of environmental safety, cost, efficiency, effectiveness, and sustainability. Researchers are also exploring the integration of microbial processes with electrochemical redox methods, namely bioelectrochemical systems (BES), as a promising technology for pollutant treatment/removal (Huang et al., 2011). BES technology has attracted increasing attention due to its high efficiency, low cost, environmental safety, room temperature operation, biocompatible materials, and the ability to recover valuable by-products (e.g., electricity, fuel, and chemicals) (Pant et al., 2012; Nazari et al., 2020). The advent of high-throughput genome sequencing and omics tools/methods has provided a wealth of new information on the genetic regulation, proteomics, and fluxomics of the reactions of various degrader microorganisms. Combining these tools with systems biology has further enhanced our understanding of the selection and fine-tuning of target catabolic pathways in microorganisms (i.e., metabolic design) to achieve efficient and effective biodegradation. To design effective bioremediation strategies using suitable candidate microorganisms, we need to understand the biochemical potential, metabolic diversity, genetic composition, and ecology (autoecology/synecology) of microorganisms.
       Fig. 1. Sources and pathways of low-molecular PAHs through various environmental environments and various factors affecting biota. Dashed lines represent interactions between ecosystem elements.
       In this review, we have attempted to summarize the data on degradation of simple PAHs such as naphthalene and substituted naphthalenes by various bacterial isolates covering metabolic pathways and diversity, enzymes involved in degradation, gene composition/content and diversity, cellular responses and various aspects of bioremediation. Understanding the biochemical and molecular levels will help in identifying suitable host strains and their further genetic engineering for effective bioremediation of such priority pollutants. This will help in developing strategies for the establishment of site-specific bacterial consortia for effective bioremediation.
       The presence of a large number of toxic and hazardous aromatic compounds (satisfying the Huckel rule 4n + 2π electrons, n = 1, 2, 3, …) poses a serious threat to various environmental media such as air, soil, sediments, and surface and groundwater (Puglisi et al., 2007). These compounds have single benzene rings (monocyclic) or multiple benzene rings (polycyclic) arranged in linear, angular or cluster form and exhibit stability (stability/instability) in the environment due to high negative resonance energy and inertness (inertness), which can be explained by their hydrophobicity and reduced state. When the aromatic ring is further replaced by methyl (-CH3), carboxyl (-COOH), hydroxyl (-OH), or sulfonate (-HSO3) groups, it becomes more stable, has a stronger affinity for macromolecules, and is bioaccumulative in biological systems (Seo et al., 2009; Phale et al., 2020). Some low molecular weight polycyclic aromatic hydrocarbons (LMWAHs), such as naphthalene and its derivatives [methylnaphthalene, naphthoic acid, naphthalenesulfonate, and 1-naphthyl N-methylcarbamate (carbaryl)], have been included in the list of priority organic pollutants by the US Environmental Protection Agency as genotoxic, mutagenic, and/or carcinogenic (Cerniglia, 1984). Release of this class of NM-PAHs into the environment may result in bioaccumulation of these compounds at all levels of the food chain, thereby affecting the health of ecosystems (Binkova et al., 2000; Srogi, 2007; Quinn et al., 2009).
       The sources and pathways of PAHs to biota are primarily through migration and interactions between different ecosystem components such as soil, groundwater, surface water, crops and the atmosphere (Arey and Atkinson, 2003). Figure 1 shows the interactions and distribution of different low molecular weight PAHs in ecosystems and their pathways to biota/human exposure. PAHs are deposited on surfaces as a result of air pollution and through the migration (drift) of vehicle emissions, industrial exhaust gases (coal gasification, combustion and coke production) and their deposition. Industrial activities such as the manufacture of synthetic textiles, dyes and paints; wood preservation; rubber processing; cement manufacturing activities; pesticide production; and agricultural applications are major sources of PAHs in terrestrial and aquatic systems (Bamforth and Singleton, 2005; Wick et al., 2011). Studies have shown that soils in suburban and urban areas, near highways, and in large cities are more susceptible to polycyclic aromatic hydrocarbons (PAHs) due to emissions from power plants, residential heating, air and road traffic loads, and construction activities (Suman et al., 2016). (2008) showed that PAHs in soil near roads in New Orleans, Louisiana, USA were as high as 7189 μg/kg, whereas in open space, they were only 2404 μg/kg. Similarly, PAH levels as high as 300 μg/kg have been reported in areas near coal gasification sites in several US cities (Kanaly and Harayama, 2000; Bamforth and Singleton, 2005). Soils from various Indian cities such as Delhi (Sharma et al., 2008), Agra (Dubey et al., 2014), Mumbai (Kulkarni and Venkataraman, 2000) and Visakhapatnam (Kulkarni et al., 2014) have been reported to contain high concentrations of PAHs. Aromatic compounds are more easily adsorbed onto soil particles, organic matter and clay minerals, thus becoming major carbon sinks in ecosystems (Srogi, 2007; Peng et al., 2008). The major sources of PAHs in aquatic ecosystems are precipitation (wet/dry precipitation and water vapour), urban runoff, wastewater discharge, groundwater recharge etc. (Srogi, 2007). It is estimated that about 80% of PAHs in marine ecosystems are derived from precipitation, sedimentation, and waste discharge (Motelay-Massei et al., 2006; Srogi, 2007). Higher concentrations of PAHs in surface water or leachate from solid waste disposal sites eventually leak into groundwater, posing a major public health threat since more than 70% of the population in South and Southeast Asia drinks groundwater (Duttagupta et al., 2019). A recent study by Duttagupta et al. (2020) of river (32) and groundwater (235) analyses from West Bengal, India, found that an estimated 53% of urban residents and 44% of rural residents (totaling 20 million residents) may be exposed to naphthalene (4.9–10.6 μg/L) and its derivatives. Differential land use patterns and increased groundwater extraction are considered to be the main factors controlling the vertical transport (advection) of low molecular weight PAHs in the subsurface. Agricultural runoff, municipal and industrial wastewater discharges, and solid waste/garbage discharges have been found to be affected by PAHs in river basins and subsurface sediments. Atmospheric precipitation further aggravates PAH pollution. High concentrations of PAHs and their alkyl derivatives (51 in total) have been reported in rivers/watersheds worldwide, such as the Fraser River, Louan River, Denso River, Missouri River, Anacostia River, Ebro River, and Delaware River (Yunker et al., 2002; Motelay-Massei et al., 2006; Li et al., 2010; Amoako et al., 2011; Kim et al., 2018). In the Ganges River basin sediments, naphthalene and phenanthrene were found to be the most significant (detected in 70% of samples) (Duttagupta et al., 2019). Moreover, studies have shown that chlorination of drinking water can lead to the formation of more toxic oxygenated and chlorinated PAHs (Manoli and Samara, 1999). PAHs accumulate in cereals, fruits and vegetables as a result of uptake by plants from contaminated soils, groundwater and precipitation (Fismes et al., 2002). Many aquatic organisms such as fish, mussels, clams and shrimp are contaminated with PAHs through the consumption of contaminated food and seawater, as well as through tissues and skin (Mackay and Fraser, 2000). Cooking/processing methods such as grilling, roasting, smoking, frying, drying, baking and charcoal cooking can also lead to significant amounts of PAHs in food. This largely depends on the choice of smoking material, phenolic/aromatic hydrocarbon content, cooking procedure, heater type, moisture content, oxygen supply and combustion temperature (Guillén et al., 2000; Gomes et al., 2013). Polycyclic aromatic hydrocarbons (PAHs) have also been detected in milk at varying concentrations (0.75–2.1 mg/L) (Girelli et al., 2014). The accumulation of these PAHs in food also depends on the physicochemical properties of food, while their toxic effects are related to physiological functions, metabolic activity, absorption, distribution and body distribution (Mechini et al., 2011).
       The toxicity and harmful effects of polycyclic aromatic hydrocarbons (PAHs) have been known for a long time (Cherniglia, 1984). Low molecular weight polycyclic aromatic hydrocarbons (LMW-PAHs) (two to three rings) can covalently bind to various macromolecules such as DNA, RNA and proteins and are carcinogenic (Santarelli et al., 2008). Due to their hydrophobic nature, they are separated by lipid membranes. In humans, cytochrome P450 monooxygenases oxidize PAHs to epoxides, some of which are highly reactive (e.g., baediol epoxide) and can lead to the transformation of normal cells into malignant ones (Marston et al., 2001). In addition, the transformation products of PAHs such as quinones, phenols, epoxides, diols, etc. are more toxic than the parent compounds. Some PAHs and their metabolic intermediates can affect hormones and various enzymes in metabolism, thereby adversely affecting growth, the central nervous system, the reproductive and immune systems (Swetha and Phale, 2005; Vamsee-Krishna et al., 2006; Oostingh et al., 2008). Short-term exposure to low molecular weight PAHs has been reported to cause impaired lung function and thrombosis in asthmatics and to increase the risk of skin, lung, bladder and gastrointestinal cancers (Olsson et al., 2010; Diggs et al., 2011). Animal studies have also shown that PAH exposure can have adverse effects on reproductive function and development and can cause cataracts, kidney and liver damage, and jaundice. Various PAH biotransformation products such as diols, epoxides, quinones and free radicals (cations) have been shown to form DNA adducts. Stable adducts have been shown to alter the DNA replication machinery, whereas unstable adducts can depurinate DNA (mainly to adenine and sometimes to guanine); both can generate errors that lead to mutations (Schweigert et al. 2001). Additionally, quinones (benzo-/pan-) can generate reactive oxygen species (ROS), causing fatal damage to DNA and other macromolecules, thereby affecting tissue function/viability (Ewa and Danuta 2017). Chronic exposure to low concentrations of pyrene, biphenyl and naphthalene has been reported to cause cancer in experimental animals (Diggs et al. 2012). Due to their lethal toxicity, cleanup/removal of these PAHs from affected/contaminated sites is a priority.
       Various physical and chemical methods have been used to remove PAHs from contaminated sites/environments. Processes such as incineration, dechlorination, UV oxidation, fixation, and solvent extraction have many disadvantages, including the formation of toxic by-products, process complexity, safety and regulatory issues, low efficiency, and high cost. However, microbial biodegradation (called bioremediation) is a promising alternative approach that involves the use of microorganisms in the form of pure cultures or colonies. Compared with physical and chemical methods, this process is environmentally friendly, non-invasive, cost-effective, and sustainable. Bioremediation can be carried out at the affected site (in situ) or at a specially prepared site (ex situ) and is therefore considered a more sustainable remediation method than traditional physical and chemical methods (Juhasz and Naidu, 2000; Andreoni and Gianfreda, 2007; Megharaj et al., 2011; Phale et al., 2020; Sarkar et al., 2020).
       Understanding the microbial metabolic steps involved in the degradation of aromatic pollutants has enormous scientific and economic implications for ecological and environmental sustainability. An estimated 2.1×1018 grams of carbon (C) is stored in sediments and organic compounds (i.e., oil, natural gas, and coal, i.e., fossil fuels) worldwide, making a significant contribution to the global carbon cycle. However, rapid industrialization, fossil fuel extraction, and human activities are depleting these lithospheric carbon reservoirs, releasing an estimated 5.5×1015 g of organic carbon (as pollutants) into the atmosphere annually (Gonzalez-Gaya et al., 2019). Most of this organic carbon enters terrestrial and marine ecosystems through sedimentation, transport, and runoff. In addition, new synthetic pollutants derived from fossil fuels, such as plastics, plasticizers and plastic stabilizers (phthalates and their isomers), seriously pollute marine, soil and aquatic ecosystems and their biota, thereby exacerbating global climate risks. Various types of microplastics, nanoplastics, plastic fragments and their toxic monomer products derived from polyethylene terephthalate (PET) have accumulated in the Pacific Ocean between North America and Southeast Asia, forming the “Great Pacific Garbage Patch”, harming marine life (Newell et al., 2020). Scientific studies have proven that it is not possible to remove such pollutants/waste by any physical or chemical methods. In this context, the most useful microorganisms are those capable of oxidatively metabolizing pollutants into carbon dioxide, chemical energy and other non-toxic by-products that eventually enter other nutrient cycling processes (H, O, N, S, P, Fe, etc.). Thus, understanding the microbial ecophysiology of aromatic pollutant mineralization and its environmental control is crucial for assessing the microbial carbon cycle, net carbon budget and future climate risks. Given the urgent need to remove such compounds from the environment, various eco-industries focused on clean technologies have emerged. Alternatively, valorization of industrial waste/waste chemicals accumulated in ecosystems (i.e. waste to wealth approach) is considered as one of the pillars of circular economy and sustainable development goals (Close et al., 2012). Therefore, understanding the metabolic, enzymatic and genetic aspects of these potential degradation candidates is of utmost importance for the effective removal and bioremediation of such aromatic pollutants.
       Among the many aromatic pollutants, we pay special attention to low-molecular-weight PAHs such as naphthalene and substituted naphthalenes. These compounds are major components of petroleum-derived fuels, textile dyes, consumer products, pesticides (mothballs and insect repellents), plasticizers and tannins and are therefore widespread in many ecosystems (Preuss et al., 2003). Recent reports highlight the accumulation of naphthalene concentrations in aquifer sediments, groundwater and subsurface soils, vadose zones and river beds, suggesting its bioaccumulation in the environment (Duttagupta et al., 2019, 2020). Table 2 summarizes the physicochemical properties, applications and health effects of naphthalene and its derivatives. Compared with other high-molecular-weight PAHs, naphthalene and its derivatives are less hydrophobic, more water-soluble and widely distributed in ecosystems, so they are often used as model substrates to study the metabolism, genetics and metabolic diversity of PAHs. A large number of microorganisms are able to metabolize naphthalene and its derivatives, and comprehensive information is available on their metabolic pathways, enzymes and regulatory features (Mallick et al., 2011; Phale et al., 2019, 2020). In addition, naphthalene and its derivatives are designated as prototype compounds for environmental pollution assessment due to their high abundance and bioavailability. The US Environmental Protection Agency estimates that average levels of naphthalene are 5.19 μg per cubic meter from cigarette smoke, primarily from incomplete combustion, and 7.8 to 46 μg from sidestream smoke, while exposure to creosote and naphthalene is 100 to 10,000 times higher (Preuss et al. 2003). Naphthalene in particular has been found to have species-, region-, and sex-specific respiratory toxicity and carcinogenicity. Based on animal studies, the International Agency for Research on Cancer (IARC) has classified naphthalene as a “possible human carcinogen” (Group 2B)1. Exposure to substituted naphthalenes, primarily by inhalation or parenteral (oral) administration, causes lung tissue injury and increases the incidence of lung tumors in rats and mice (National Toxicology Program 2). Acute effects include nausea, vomiting, abdominal pain, diarrhea, headache, confusion, profuse sweating, fever, tachycardia, etc. On the other hand, the broad-spectrum carbamate insecticide carbaryl (1-naphthyl N-methylcarbamate) has been reported to be toxic to aquatic invertebrates, amphibians, honey bees and humans and has been shown to inhibit acetylcholinesterase causing paralysis (Smulders et al., 2003; Bulen and Distel, 2011). Therefore, understanding the mechanisms of microbial degradation, genetic regulation, enzymatic and cellular reactions is crucial for developing bioremediation strategies in contaminated environments.
       Table 2. Detailed information on the physicochemical properties, uses, identification methods and associated diseases of naphthalene and its derivatives.
       In polluted niches, hydrophobic and lipophilic aromatic pollutants can cause a variety of cellular effects on the environmental microbiome (community), such as changes in membrane fluidity, membrane permeability, lipid bilayer swelling, disruption of energy transfer (electron transport chain/proton motive force), and activity of membrane-associated proteins (Sikkema et al., 1995). In addition, some soluble intermediates such as catechols and quinones generate reactive oxygen species (ROS) and form adducts with DNA and proteins (Penning et al., 1999). Thus, the abundance of such compounds in ecosystems exerts selective pressure on microbial communities to become efficient degraders at various physiological levels, including uptake/transport, intracellular transformation, assimilation/utilization, and compartmentalization.
       A search of the Ribosomal Database Project-II (RDP-II) revealed that a total of 926 bacterial species were isolated from media or enrichment cultures contaminated with naphthalene or its derivatives. The Proteobacteria group had the highest number of representatives (n = 755), followed by Firmicutes (52), Bacteroidetes (43), Actinobacteria (39), Tenericutes (10), and unclassified bacteria (8) (Figure 2). Representatives of γ-Proteobacteria (Pseudomonadales and Xanthomonadales) dominated all Gram-negative groups with high G+C content (54%), while Clostridiales and Bacillales (30%) were Gram-positive groups with low G+C content. Pseudomonas (the highest number, 338 species) were reported to be able to degrade naphthalene and its methyl derivatives in various polluted ecosystems (coal tar, petroleum, crude oil, sludge, oil spills, wastewater, organic waste and landfills) as well as in intact ecosystems (soil, rivers, sediments and groundwater) (Figure 2). Moreover, enrichment studies and metagenomic analysis of some of these regions revealed that uncultured Legionella and Clostridium species may have degradative capacity, indicating the need to culture these bacteria to study new pathways and metabolic diversity.
       Fig. 2. Taxonomic diversity and ecological distribution of bacterial representatives in environments contaminated with naphthalene and naphthalene derivatives.
       Among the various aromatic hydrocarbon-degrading microorganisms, most are capable of degrading naphthalene as the sole source of carbon and energy. The sequence of events involved in naphthalene metabolism has been described for Pseudomonas sp. (strains: NCIB 9816-4, G7, AK-5, PMD-1 and CSV86), Pseudomonas stutzeri AN10, Pseudomonas fluorescens PC20 and other strains (ND6 and AS1) (Mahajan et al., 1994; Resnick et al., 1996; Annweiler et al., 2000; Basu et al., 2003; Dennis and Zylstra, 2004; Sota et al., 2006; Metabolism is initiated by a multicomponent dioxygenase [naphthalene dioxygenase (NDO), a ring hydroxylating dioxygenase] that catalyzes the oxidation of one of the aromatic rings of naphthalene using molecular oxygen as the other substrate, converting naphthalene to cis-naphthalenediol (Figure 3). Cis-dihydrodiol is converted to 1,2-dihydroxynaphthalene by a dehydrogenase. A ring-cleaving dioxygenase, 1,2-dihydroxynaphthalene dioxygenase (12DHNDO), converts 1,2-dihydroxynaphthalene to 2-hydroxychromene-2-carboxylic acid. Enzymatic cis-trans isomerization produces trans-o-hydroxybenzylidenepyruvate, which is cleaved by hydratase aldolase to salicylic aldehyde and pyruvate. The organic acid pyruvate was the first C3 compound derived from the naphthalene carbon skeleton and directed into the central carbon pathway. In addition, NAD+-dependent salicylaldehyde dehydrogenase converts salicylaldehyde to salicylic acid. Metabolism at this stage is called the “upper pathway” of naphthalene degradation. This pathway is very common in most naphthalene-degrading bacteria. However, there are a few exceptions; for example, in the thermophilic Bacillus hamburgii 2, naphthalene degradation is initiated by naphthalene 2,3-dioxygenase to form 2,3-dihydroxynaphthalene (Annweiler et al., 2000).
       Figure 3. Pathways of naphthalene, methylnaphthalene, naphthoic acid, and carbaryl degradation. Circled numbers represent enzymes responsible for the sequential conversion of naphthalene and its derivatives into subsequent products. 1 — naphthalene dioxygenase (NDO); 2, cis-dihydrodiol dehydrogenase; 3, 1,2-dihydroxynaphthalene dioxygenase; 4, 2-hydroxychromene-2-carboxylic acid isomerase; 5, trans-O-hydroxybenzylidenepyruvate hydratase aldolase; 6, salicylaldehyde dehydrogenase; 7, salicylate 1-hydroxylase; 8, catechol 2,3-dioxygenase (C23DO); 9, 2-hydroxymuconate semialdehyde dehydrogenase; 10, 2-oxopent-4-enoate hydratase; 11, 4-hydroxy-2-oxopentanoate aldolase; 12, acetaldehyde dehydrogenase; 13, catechol-1,2-dioxygenase (C12DO); 14, muconate cycloisomerase; 15, muconolactone delta-isomerase; 16, β-ketoadipatenollactone hydrolase; 17, β-ketoadipate succinyl-CoA transferase; 18, β-ketoadipate-CoA thiolase; 19, succinyl-CoA: acetyl-CoA succinyltransferase; 20, salicylate 5-hydroxylase; 21 – gentisate 1,2-dioxygenase (GDO); 22, maleylpyruvate isomerase; 23, fumarylpyruvate hydrolase; 24, methylnaphthalene hydroxylase (NDO); 25, hydroxymethylnaphthalene dehydrogenase; 26, naphthaldehyde dehydrogenase; 27, 3-formylsalicylic acid oxidase; 28, hydroxyisophthalate decarboxylase; 29, carbaryl hydrolase (CH); 30, 1-naphthol-2-hydroxylase.
       Depending on the organism and its genetic makeup, the resulting salicylic acid is further metabolized either via the catechol pathway using salicylate 1-hydroxylase (S1H) or via the gentisate pathway using salicylate 5-hydroxylase (S5H) (Figure 3). Since salicylic acid is the major intermediate in naphthalene metabolism (upper pathway), the steps from salicylic acid to the TCA intermediate are often referred to as the lower pathway, and the genes are organized into a single operon. It is common to see that the genes in the upper pathway operon (nah) and the lower pathway operon (sal) are regulated by common regulatory factors; for example, NahR and salicylic acid act as inducers, allowing both operons to completely metabolize naphthalene (Phale et al., 2019, 2020).
       In addition, catechol is cyclically cleaved to 2-hydroxymuconate semialdehyde via the meta pathway by catechol 2,3-dioxygenase (C23DO) (Yen et al., 1988) and further hydrolyzed by 2-hydroxymuconate semialdehyde hydrolase to form 2-hydroxypent-2,4-dienoic acid. 2-hydroxypent-2,4-dienoate is then converted to pyruvate and acetaldehyde by a hydratase (2-oxopent-4-enoate hydratase) and an aldolase (4-hydroxy-2-oxopentanoate aldolase) and then enters the central carbon pathway (Figure 3). Alternatively, catechol is cyclically cleaved to cis,cis-muconate via the ortho pathway by catechol 1,2-oxygenase (C12DO). Muconate cycloisomerase, muconolactone isomerase, and β-ketoadipate-nollactone hydrolase convert cis,cis-muconate to 3-oxoadipate, which enters the central carbon pathway via succinyl-CoA and acetyl-CoA (Nozaki et al., 1968) (Figure 3).
       In the gentisate (2,5-dihydroxybenzoate) pathway, the aromatic ring is cleaved by gentisate 1,2-dioxygenase (GDO) to form maleylpyruvate. This product can be directly hydrolyzed to pyruvate and malate, or it can be isomerized to form fumarylpyruvate, which can then be hydrolyzed to pyruvate and fumarate (Larkin and Day, 1986). The choice of the alternative pathway has been observed in both Gram-negative and Gram-positive bacteria at the biochemical and genetic levels (Morawski et al., 1997; Whyte et al., 1997). Gram-negative bacteria (Pseudomonas) prefer to use salicylic acid, which is an inducer of naphthalene metabolism, decarboxylating it to catechol using salicylate 1-hydroxylase (Gibson and Subramanian, 1984). On the other hand, in Gram-positive bacteria (Rhodococcus), salicylate 5-hydroxylase converts salicylic acid to gentisic acid, whereas salicylic acid has no inductive effect on the transcription of naphthalene genes (Grund et al., 1992) (Figure 3).
       It has been reported that species such as Pseudomonas CSV86, Oceanobacterium NCE312, Marinhomonas naphthotrophicus, Sphingomonas paucimobilis 2322, Vibrio cyclotrophus, Pseudomonas fluorescens LP6a, Pseudomonas and Mycobacterium species can degrade monomethylnaphthalene or dimethylnaphthalene (Dean-Raymond and Bartha, 1975; Cane and Williams, 1982; Mahajan et al., 1994; Dutta et al., 1998; Hedlund et al., 1999). Among them, the 1-methylnaphthalene and 2-methylnaphthalene degradation pathway of Pseudomonas sp. CSV86 has been clearly studied at the biochemical and enzymatic levels (Mahajan et al., 1994). 1-Methylnaphthalene is metabolized via two pathways. First, the aromatic ring is hydroxylated (the unsubstituted ring of methylnaphthalene) to form cis-1,2-dihydroxy-1,2-dihydro-8-methylnaphthalene, which is further oxidized to methyl salicylate and methylcatechol, and then enters the central carbon pathway after ring cleavage (Figure 3). This pathway is called the “carbon source pathway”. In the second “detoxification pathway”, the methyl group can be hydroxylated by NDO to form 1-hydroxymethylnaphthalene, which is further oxidized to 1-naphthoic acid and excreted into the culture medium as a dead-end product. Studies have shown that strain CSV86 is unable to grow on 1- and 2-naphthoic acid as the sole carbon and energy source, confirming its detoxification pathway (Mahajan et al., 1994; Basu et al., 2003). In 2-methylnaphthalene, the methyl group undergoes hydroxylation by hydroxylase to form 2-hydroxymethylnaphthalene. In addition, the unsubstituted ring of the naphthalene ring undergoes ring hydroxylation to form a dihydrodiol, which is oxidized to 4-hydroxymethylcatechol in a series of enzyme-catalyzed reactions and enters the central carbon pathway via the meta-ring cleavage pathway. Similarly, S. paucimobilis 2322 was reported to utilize NDO to hydroxylate 2-methylnaphthalene, which is further oxidized to form methyl salicylate and methylcatechol (Dutta et al., 1998).
       Naphthoic acids (substituted/unsubstituted) are detoxification/biotransformation by-products formed during the degradation of methylnaphthalene, phenanthrene and anthracene and released into the spent culture medium. It has been reported that the soil isolate Stenotrophomonas maltophilia CSV89 is able to metabolize 1-naphthoic acid as a carbon source (Phale et al., 1995). Metabolism begins with dihydroxylation of the aromatic ring to form 1,2-dihydroxy-8-carboxynaphthalene. The resulting diol is oxidized to catechol via 2-hydroxy-3-carboxybenzylidenepyruvate, 3-formylsalicylic acid, 2-hydroxyisophthalic acid and salicylic acid and enters the central carbon pathway via the meta-ring cleavage pathway (Figure 3).
       Carbaryl is a naphthyl carbamate pesticide. Since the Green Revolution in India in the 1970s, the use of chemical fertilizers and pesticides has led to an increase in polycyclic aromatic hydrocarbon (PAH) emissions from agricultural non-point sources (Pingali, 2012; Duttagupta et al., 2020). An estimated 55% (85,722,000 hectares) of total cropland in India is treated with chemical pesticides. Over the last five years (2015–2020), the Indian agriculture sector has used an average of 55,000 to 60,000 tonnes of pesticides annually (Department of Cooperatives and Farmers Welfare, Ministry of Agriculture, Government of India, August 2020). In the northern and central Gangetic plains (the states with the highest population and population density), the use of pesticides on crops is widespread, with insecticides predominating. Carbaryl (1-naphthyl-N-methylcarbamate) is a broad-spectrum, moderately to highly toxic carbamate insecticide used in Indian agriculture at an average rate of 100–110 tonnes. It is commonly sold under the trade name Sevin and is used to control insects (aphids, fire ants, fleas, mites, spiders and many other outdoor pests) affecting a variety of crops (maize, soybean, cotton, fruits and vegetables). Some microorganisms such as Pseudomonas (NCIB 12042, 12043, C4, C5, C6, C7, Pseudomonas putida XWY-1), Rhodococcus (NCIB 12038), Sphingobacterium spp. (CF06), Burkholderia (C3), Micrococcus and Arthrobacter can also be used to control other pests. It has been reported that RC100 can degrade carbaryl (Larkin and Day, 1986; Chapalamadugu and Chaudhry, 1991; Hayatsu et al., 1999; Swetha and Phale, 2005; Trivedi et al., 2017). The degradation pathway of carbaryl has been extensively studied at biochemical, enzymatic and genetic levels in soil isolates of Pseudomonas sp. Strains C4, C5 and C6 (Swetha and Phale, 2005; Trivedi et al., 2016) (Fig. 3). The metabolic pathway starts with the hydrolysis of the ester bond by carbaryl hydrolase (CH) to form 1-naphthol, methylamine and carbon dioxide. 1-naphthol is then converted to 1,2-dihydroxynaphthalene by 1-naphthol hydroxylase (1-NH), which is further metabolized via the central carbon pathway via salicylate and gentisate. Some carbaryl-degrading bacteria have been reported to metabolize it to salicylic acid via cleavage of the catechol ortho ring (Larkin and Day, 1986; Chapalamadugu and Chaudhry, 1991). Notably, naphthalene-degrading bacteria primarily metabolize salicylic acid via catechol, whereas carbaryl-degrading bacteria prefer to metabolize salicylic acid via the gentisate pathway.
       Naphthalenesulfonic acid/disulfonic acid and naphthylaminesulfonic acid derivatives can be used as intermediates in the production of azo dyes, wetting agents, dispersants, etc. Although these compounds have low toxicity to humans, cytotoxicity assessments have shown that they are lethal to fish, daphnia and algae (Greim et al., 1994). Representatives of the genus Pseudomonas (strains A3, C22) have been reported to initiate metabolism by double hydroxylation of the aromatic ring containing the sulfonic acid group to form a dihydrodiol, which is further converted to 1,2-dihydroxynaphthalene by spontaneous cleavage of the sulfite group (Brilon et al., 1981). The resulting 1,2-dihydroxynaphthalene is catabolized via the classical naphthalene pathway, i.e., the catechol or gentisate pathway (Figure 4). It has been shown that aminonaphthalenesulfonic acid and hydroxynaphthalenesulfonic acid can be completely degraded by mixed bacterial consortia with complementary catabolic pathways (Nortemann et al., 1986). It has been shown that one member of the consortium desulfurizes aminonaphthalenesulfonic acid or hydroxynaphthalenesulfonic acid by 1,2-dioxygenation, while aminosalicylate or hydroxysalicylate is released into the culture medium as a dead-end metabolite and is subsequently taken up by other members of the consortium. Naphthalenedisulfonic acid is relatively polar but poorly biodegradable and can therefore be metabolized via different pathways. The first desulfurization occurs during regioselective dihydroxylation of the aromatic ring and the sulfonic acid group; the second desulfurization occurs during hydroxylation of 5-sulfosalicylic acid by salicylic acid 5-hydroxylase to form gentisic acid, which enters the central carbon pathway (Brilon et al., 1981) (Figure 4). The enzymes responsible for naphthalene degradation are also responsible for naphthalene sulfonate metabolism (Brilon et al., 1981; Keck et al., 2006).
       Figure 4. Metabolic pathways for naphthalene sulfonate degradation. The numbers inside the circles represent the enzymes responsible for naphthyl sulfonate metabolism, similar/identical to the enzymes described in FIG. 3.
       Low molecular weight PAHs (LMW-PAHs) are reducible, hydrophobic and poorly soluble, and therefore not susceptible to natural breakdown/degradation. However, aerobic microorganisms are able to oxidize it by absorbing molecular oxygen (O2). These enzymes mainly belong to the class of oxidoreductases and can perform various reactions such as aromatic ring hydroxylation (mono- or dihydroxylation), dehydrogenation and aromatic ring cleavage. The products obtained from these reactions are in a higher oxidation state and are more easily metabolized through the central carbon pathway (Phale et al., 2020). The enzymes in the degradation pathway have been reported to be inducible. The activity of these enzymes is very low or negligible when cells are grown on simple carbon sources such as glucose or organic acids. Table 3 summarizes the various enzymes (oxygenases, hydrolases, dehydrogenases, oxidases, etc.) involved in the metabolism of naphthalene and its derivatives.
       Table 3. Biochemical characteristics of enzymes responsible for the degradation of naphthalene and its derivatives.
       Radioisotope studies (18O2) have shown that the incorporation of molecular O2 into aromatic rings by oxygenases is the most important step in activating further biodegradation of a compound (Hayaishi et al., 1955; Mason et al., 1955). The incorporation of one oxygen atom (O) from molecular oxygen (O2) into the substrate is initiated by either endogenous or exogenous monooxygenases (also called hydroxylases). Another oxygen atom is reduced to water. Exogenous monooxygenases reduce flavin with NADH or NADPH, whereas in endomonooxygenases flavin is reduced by the substrate. The position of hydroxylation results in diversity in product formation. For example, salicylate 1-hydroxylase hydroxylates salicylic acid at the C1 position, forming catechol. On the other hand, the multicomponent salicylate 5-hydroxylase (containing reductase, ferredoxin, and oxygenase subunits) hydroxylates salicylic acid at the C5 position, forming gentisic acid (Yamamoto et al., 1965).
       Dioxygenases incorporate two O2 atoms into the substrate. Depending on the products formed, they are divided into ring hydroxylating dioxygenases and ring cleaving dioxygenases. Ring hydroxylating dioxygenases convert aromatic substrates into cis-dihydrodiols (e.g., naphthalene) and are widespread among bacteria. To date, it has been shown that organisms containing ring hydroxylating dioxygenases are capable of growing on various aromatic carbon sources, and these enzymes are classified as NDO (naphthalene), toluene dioxygenase (TDO, toluene), and biphenyl dioxygenase (BPDO, biphenyl). Both NDO and BPDO can catalyze the double oxidation and side chain hydroxylation of various polycyclic aromatic hydrocarbons (toluene, nitrotoluene, xylene, ethylbenzene, naphthalene, biphenyl, fluorene, indole, methylnaphthalene, naphthalenesulfonate, phenanthrene, anthracene, acetophenone, etc.) (Boyd and Sheldrake, 1998; Phale et al., 2020). NDO is a multicomponent system consisting of an oxidoreductase, a ferredoxin, and an active site-containing oxygenase component (Gibson and Subramanian, 1984; Resnick et al., 1996). The catalytic unit of NDO consists of a large α subunit and a small β subunit arranged in an α3β3 configuration. NDO belongs to a large family of oxygenases and its α-subunit contains a Rieske site [2Fe-2S] and a mononuclear non-heme iron, which determines the substrate specificity of NDO (Parales et al., 1998). Typically, in one catalytic cycle, two electrons from the reduction of pyridine nucleotide are transferred to the Fe(II) ion in the active site via a reductase, a ferredoxin and a Rieske site. The reducing equivalents activate molecular oxygen, which is a prerequisite for substrate dihydroxylation (Ferraro et al., 2005). To date, only a few NDOs have been purified and characterized in detail from different strains and the genetic control of the pathways involved in naphthalene degradation has been studied in detail (Resnick et al., 1996; Parales et al., 1998; Karlsson et al., 2003). Ring-cleaving dioxygenases (endo- or ortho-ring-cleaving enzymes and exodiol- or meta-ring-cleaving enzymes) act on hydroxylated aromatic compounds. For example, the ortho-ring-cleaving dioxygenase is catechol-1,2-dioxygenase, whereas the meta-ring-cleaving dioxygenase is catechol-2,3-dioxygenase (Kojima et al., 1961; Nozaki et al., 1968). In addition to various oxygenases, there are also various dehydrogenases responsible for the dehydrogenation of aromatic dihydrodiols, alcohols and aldehydes and using NAD+/NADP+ as electron acceptors, which are some of the important enzymes involved in metabolism (Gibson and Subramanian, 1984; Shaw and Harayama, 1990; Fahle et al., 2020).
       Enzymes such as hydrolases (esterases, amidases) are a second important class of enzymes that use water to cleave covalent bonds and exhibit broad substrate specificity. Carbaryl hydrolase and other hydrolases are considered to be components of the periplasm (transmembrane) in members of Gram-negative bacteria (Kamini et al., 2018). Carbaryl has both an amide and an ester linkage; therefore, it can be hydrolyzed by either esterase or amidase to form 1-naphthol. Carbaryl in Rhizobium rhizobium strain AC10023 and Arthrobacter strain RC100 has been reported to function as an esterase and amidase, respectively. Carbaryl in Arthrobacter strain RC100 also functions as an amidase. RC100 has been shown to hydrolyze four N-methylcarbamate class insecticides such as carbaryl, methomyl, mefenamic acid and XMC (Hayaatsu et al., 2001). It was reported that CH in Pseudomonas sp. C5pp can act on carbaryl (100% activity) and 1-naphthyl acetate (36% activity), but not on 1-naphthylacetamide, indicating that it is an esterase (Trivedi et al., 2016).
       Biochemical studies, enzyme regulation patterns, and genetic analysis have shown that the naphthalene degradation genes consist of two inducible regulatory units or “operons”: nah (the “upstream pathway”, converting naphthalene to salicylic acid) and sal (the “downstream pathway”, converting salicylic acid to the central carbon pathway via catechol). Salicylic acid and its analogues can act as inducers (Shamsuzzaman and Barnsley, 1974). In the presence of glucose or organic acids, the operon is repressed. Figure 5 shows the complete genetic organization of naphthalene degradation (in operon form). Several named variants/forms of the nah gene (ndo/pah/dox) have been described and found to have high sequence homology (90%) among all Pseudomonas species (Abbasian et al., 2016). The genes of the naphthalene upstream pathway were generally arranged in a consensus order as shown in Figure 5A. Another gene, nahQ, was also reported to be involved in naphthalene metabolism and was usually located between nahC and nahE, but its actual function remains to be elucidated. Similarly, the nahY gene, responsible for naphthalene-sensitive chemotaxis, was found at the distal end of the nah operon in some members. In Ralstonia sp., the U2 gene encoding glutathione S-transferase (gsh) was found to be located between nahAa and nahAb but did not affect the naphthalene utilization characteristics (Zylstra et al., 1997).
       Figure 5. Genetic organization and diversity observed during naphthalene degradation among bacterial species; (A) Upper naphthalene pathway, metabolism of naphthalene to salicylic acid; (B) Lower naphthalene pathway, salicylic acid via catechol to the central carbon pathway; (C) salicylic acid via gentisate to the central carbon pathway.
       The “lower pathway” (sal operon) typically consists of nahGTHINLMOKJ and converts salicylate to pyruvate and acetaldehyde via the catechol metaring cleavage pathway. The nahG gene (encoding salicylate hydroxylase) was found to be conserved at the proximal end of the operon (Fig. 5B). Compared with other naphthalene-degrading strains, in P. putida CSV86 the nah and sal operons are tandem and very closely related (about 7.5 kb). In some Gram-negative bacteria, such as Ralstonia sp. U2, Polaromonas naphthalenivorans CJ2, and P. putida AK5, naphthalene is metabolized as a central carbon metabolite via the gentisate pathway (in the form of the sgp/nag operon). The gene cassette is typically represented in the form nagAaGHAbAcAdBFCQEDJI, where nagR (encoding a LysR-type regulator) is located at the upper end (Figure 5C).
       Carbaryl enters the central carbon cycle via the metabolism of 1-naphthol, 1,2-dihydroxynaphthalene, salicylic acid, and gentisic acid (Figure 3). Based on genetic and metabolic studies, it has been proposed to divide this pathway into “upstream” (conversion of carbaryl to salicylic acid), “middle” (conversion of salicylic acid to gentisic acid), and “downstream” (conversion of gentisic acid to central carbon pathway intermediates) (Singh et al., 2013). Genomic analysis of C5pp (supercontig A, 76.3 kb) revealed that the mcbACBDEF gene is involved in the conversion of carbaryl to salicylic acid, followed by mcbIJKL in the conversion of salicylic acid to gentisic acid, and mcbOQP in the conversion of gentisic acid to central carbon intermediates (fumarate and pyruvate, Trivedi et al., 2016) (Figure 6).
       It has been reported that enzymes involved in the degradation of aromatic hydrocarbons (including naphthalene and salicylic acid) can be induced by the corresponding compounds and inhibited by simple carbon sources such as glucose or organic acids (Shingler, 2003; Phale et al., 2019, 2020). Among the various metabolic pathways of naphthalene and its derivatives, the regulatory features of naphthalene and carbaryl have been studied to some extent. For naphthalene, genes in both the upstream and downstream pathways are regulated by NahR, a LysR-type trans-acting positive regulator. It is required for the induction of the nah gene by salicylic acid and its subsequent high-level expression (Yen and Gunsalus, 1982). Furthermore, studies have shown that integrative host factor (IHF) and XylR (sigma 54-dependent transcriptional regulator) are also critical for the transcriptional activation of genes in naphthalene metabolism (Ramos et al., 1997). Studies have shown that enzymes of the catechol meta-ring opening pathway, namely catechol 2,3-dioxygenase, are induced in the presence of naphthalene and/or salicylic acid (Basu et al., 2006). Studies have shown that enzymes of the catechol ortho-ring opening pathway, namely catechol 1,2-dioxygenase, are induced in the presence of benzoic acid and cis,cis-muconate (Parsek et al., 1994; Tover et al., 2001).
       In strain C5pp, five genes, mcbG, mcbH, mcbN, mcbR and mcbS, encode regulators belonging to the LysR/TetR family of transcriptional regulators responsible for controlling carbaryl degradation. The homologous gene mcbG was found to be most closely related to the LysR-type regulator PhnS (58% amino acid identity) involved in phenanthrene metabolism in Burkholderia RP00725 (Trivedi et al., 2016). The mcbH gene was found to be involved in the intermediate pathway (conversion of salicylic acid to gentisic acid) and belongs to the LysR-type transcriptional regulator NagR/DntR/NahR in Pseudomonas and Burkholderia. Members of this family were reported to recognize salicylic acid as a specific effector molecule for the induction of degradation genes. On the other hand, three genes, mcbN, mcbR and mcbS, belonging to LysR and TetR type transcriptional regulators, were identified in the downstream pathway (gentisate-central carbon pathway metabolites).
       In prokaryotes, horizontal gene transfer processes (acquisition, exchange, or transfer) via plasmids, transposons, prophages, genomic islands, and integrative conjugative elements (ICE) are major causes of plasticity in bacterial genomes, leading to the gain or loss of specific functions/traits. It allows bacteria to rapidly adapt to different environmental conditions, providing potential adaptive metabolic advantages to the host, such as the degradation of aromatic compounds. Metabolic changes are often achieved through fine-tuning of degradation operons, their regulatory mechanisms, and enzyme specificities, which facilitates the degradation of a wider range of aromatic compounds (Nojiri et al., 2004; Phale et al., 2019, 2020). The gene cassettes for naphthalene degradation have been found to be located on a variety of mobile elements such as plasmids (conjugative and non-conjugative), transposons, genomes, ICEs, and combinations of different bacterial species (Figure 5). In Pseudomonas G7, the nah and sal operons of plasmid NAH7 are transcribed in the same orientation and are part of a defective transposon that requires transposase Tn4653 for mobilization (Sota et al., 2006). In Pseudomonas strain NCIB9816-4, the gene was found on the conjugative plasmid pDTG1 as two operons (approximately 15 kb apart) that were transcribed in opposite directions (Dennis and Zylstra, 2004). In Pseudomonas putida strain AK5, the non-conjugative plasmid pAK5 encodes the enzyme responsible for naphthalene degradation via the gentisate pathway (Izmalkova et al., 2013). In Pseudomonas strain PMD-1, the nah operon is located on the chromosome, whereas the sal operon is located on the conjugative plasmid pMWD-1 (Zuniga et al., 1981). However, in Pseudomonas stutzeri AN10, all naphthalene degradation genes (nah and sal operons) are located on the chromosome and are presumably recruited through transposition, recombination, and rearrangement events (Bosch et al., 2000). In Pseudomonas sp. CSV86, the nah and sal operons are located in the genome in the form of ICE (ICECSV86). The structure is protected by tRNAGly followed by direct repeats indicating recombination/attachment sites (attR and attL) and a phage-like integrase located at both ends of tRNAGly, thus structurally similar to the ICEclc element (ICEclcB13 in Pseudomonas knackmusii for chlorocatechol degradation). It has been reported that genes on ICE can be transferred by conjugation with an extremely low transfer frequency (10-8), thereby transferring degradation properties to the recipient (Basu and Phale, 2008; Phale et al., 2019).
       Most of the genes responsible for carbaryl degradation are located on plasmids. Arthrobacter sp. RC100 contains three plasmids (pRC1, pRC2 and pRC300) of which two conjugative plasmids, pRC1 and pRC2, encode the enzymes that convert carbaryl to gentisate. On the other hand, the enzymes involved in the conversion of gentisate to the central carbon metabolites are located on the chromosome (Hayaatsu et al., 1999). Bacteria of the genus Rhizobium. Strain AC100, used for the conversion of carbaryl to 1-naphthol, contain plasmid pAC200, which carries the cehA gene encoding CH as part of the Tnceh transposon surrounded by insertion element-like sequences (istA and istB) (Hashimoto et al., 2002). In Sphingomonas strain CF06, the carbaryl degradation gene is believed to be present in five plasmids: pCF01, pCF02, pCF03, pCF04, and pCF05. The DNA homology of these plasmids is high, indicating the existence of a gene duplication event (Feng et al., 1997). In a carbaryl-degrading symbiont composed of two Pseudomonas species, strain 50581 contains a conjugative plasmid pCD1 (50 kb) encoding the mcd carbaryl hydrolase gene, whereas the conjugative plasmid in strain 50552 encodes a 1-naphthol-degrading enzyme (Chapalamadugu and Chaudhry, 1991). In Achromobacter strain WM111, the mcd furadan hydrolase gene is located on a 100 kb plasmid (pPDL11). This gene has been shown to be present on different plasmids (100, 105, 115 or 124 kb) in different bacteria from different geographical regions (Parekh et al., 1995). In Pseudomonas sp. C5pp, all genes responsible for carbaryl degradation are located in a genome spanning 76.3 kb of sequence (Trivedi et al., 2016). Genome analysis (6.15 Mb) revealed the presence of 42 MGEs and 36 GEIs, of which 17 MGEs were located in supercontig A (76.3 kb) with an average asymmetric G+C content (54–60 mol%), suggesting possible horizontal gene transfer events (Trivedi et al., 2016). P. putida XWY-1 exhibits a similar arrangement of carbaryl-degrading genes, but these genes are located on a plasmid (Zhu et al., 2019).
       In addition to metabolic efficiency at biochemical and genomic levels, microorganisms also exhibit other properties or responses such as chemotaxis, cell surface modification properties, compartmentalization, preferential utilization, biosurfactant production, etc., which help them to more efficiently metabolize aromatic pollutants in contaminated environments (Figure 7).
       Figure 7. Different cellular response strategies of ideal aromatic hydrocarbon-degrading bacteria for efficient biodegradation of foreign pollutant compounds.
       Chemotactic responses are considered to be factors enhancing the degradation of organic pollutants in heterogeneously polluted ecosystems. (2002) demonstrated that chemotaxis of Pseudomonas sp. G7 to naphthalene increased the rate of naphthalene degradation in aquatic systems. The wild-type strain G7 degraded naphthalene much faster than a chemotaxis-deficient mutant strain. The NahY protein (538 amino acids with membrane topology) was found to be co-transcribed with the metacleavage pathway genes on the NAH7 plasmid, and like chemotaxis transducers, this protein appears to function as a chemoreceptor for naphthalene degradation (Grimm and Harwood 1997). Another study by Hansel et al. (2009) showed that the protein is chemotactic, but its degradation rate is high. (2011) demonstrated a chemotactic response of Pseudomonas (P. putida) to gaseous naphthalene, wherein gas phase diffusion resulted in a steady flow of naphthalene to the cells, which controlled the chemotactic response of the cells. The researchers exploited this chemotactic behavior to engineer microbes that would enhance the rate of degradation. Studies have shown that chemosensory pathways also regulate other cellular functions such as cell division, cell cycle regulation, and biofilm formation, thereby helping to control the rate of degradation. However, harnessing this property (chemotaxis) for efficient degradation is hampered by several bottlenecks. The major hurdles are: (a) different paralogous receptors recognize the same compounds/ligands; (b) existence of alternative receptors, i.e., energetic tropism; (c) significant sequence differences in the sensory domains of the same receptor family; and (d) lack of information on the major bacterial sensor proteins (Ortega et al., 2017; Martin-Mora et al., 2018). Sometimes, the biodegradation of aromatic hydrocarbons produces multiple metabolites/intermediates, which may be chemotactic for one group of bacteria but repulsive for others, further complicating the process. To identify the interactions of ligands (aromatic hydrocarbons) with chemical receptors, we constructed hybrid sensor proteins (PcaY, McfR, and NahY) by fusing the sensor and signaling domains of Pseudomonas putida and Escherichia coli, which target the receptors for aromatic acids, TCA intermediates, and naphthalene, respectively (Luu et al., 2019).
       Under the influence of naphthalene and other polycyclic aromatic hydrocarbons (PAHs), the structure of the bacterial membrane and the integrity of the microorganisms undergo significant changes. Studies have shown that naphthalene interferes with the interaction of the acyl chain through hydrophobic interactions, thereby increasing the swelling and fluidity of the membrane (Sikkema et al., 1995). To counteract this detrimental effect, bacteria regulate membrane fluidity by changing the ratio and fatty acid composition between iso/anteiso branched-chain fatty acids and isomerizing cis-unsaturated fatty acids into the corresponding trans-isomers (Heipieper and de Bont, 1994). In Pseudomonas stutzeri grown on naphthalene treatment, the saturated to unsaturated fatty acid ratio increased from 1.1 to 2.1, whereas in Pseudomonas JS150 this ratio increased from 7.5 to 12.0 (Mrozik et al., 2004). When grown on naphthalene, Achromobacter KAs 3–5 cells exhibited cell aggregation around naphthalene crystals and a decrease in cell surface charge (from -22.5 to -2.5 mV) accompanied by cytoplasmic condensation and vacuolization, indicating changes in cell structure and cell surface properties (Mohapatra et al., 2019). Although cellular/surface changes are directly associated with better uptake of aromatic pollutants, relevant bioengineering strategies have not been thoroughly optimized. Manipulation of cell shape has rarely been used to optimize biological processes (Volke and Nikel, 2018). Deletion of genes affecting cell division causes changes in cell morphology. Deletion of genes affecting cell division causes changes in cell morphology. In Bacillus subtilis, the cell septum protein SepF has been shown to be involved in septum formation and is required for subsequent steps of cell division, but it is not an essential gene. Deletion of genes encoding peptide glycan hydrolases in Bacillus subtilis resulted in cell elongation, increased specific growth rate, and improved enzyme production capacity (Cui et al., 2018).
       Compartmentalization of the carbaryl degradation pathway has been proposed to achieve efficient degradation of Pseudomonas strains C5pp and C7 (Kamini et al., 2018). It is proposed that carbaryl is transported into the periplasmic space through the outer membrane septum and/or through diffusible porins. CH is a periplasmic enzyme that catalyzes the hydrolysis of carbaryl to 1-naphthol, which is more stable, more hydrophobic and more toxic. CH is localized in the periplasm and has a low affinity for carbaryl, thus controlling the formation of 1-naphthol, thereby preventing its accumulation in cells and reducing its toxicity to cells (Kamini et al., 2018). The resulting 1-naphthol is transported into the cytoplasm across the inner membrane by partitioning and/or diffusion, and is then hydroxylated to 1,2-dihydroxynaphthalene by the high-affinity enzyme 1NH for further metabolism in the central carbon pathway.
       Although microorganisms have the genetic and metabolic capabilities to degrade xenobiotic carbon sources, the hierarchical structure of their utilization (i.e., preferential use of simple over complex carbon sources) is a major obstacle to biodegradation. The presence and utilization of simple carbon sources downregulates genes encoding enzymes that degrade complex/non-preferred carbon sources such as PAHs. A well-studied example is that when glucose and lactose are co-fed to Escherichia coli, glucose is utilized more efficiently than lactose (Jacob and Monod, 1965). Pseudomonas has been reported to degrade a variety of PAHs and xenobiotic compounds as carbon sources. The hierarchy of carbon source utilization in Pseudomonas is organic acids > glucose > aromatic compounds (Hylemon and Phibbs, 1972; Collier et al., 1996). However, there is an exception. Interestingly, Pseudomonas sp. CSV86 exhibits a unique hierarchical structure that preferentially utilizes aromatic hydrocarbons (benzoic acid, naphthalene, etc.) rather than glucose and co-metabolizes aromatic hydrocarbons with organic acids (Basu et al., 2006). In this bacterium, the genes for degradation and transport of aromatic hydrocarbons are not downregulated even in the presence of a second carbon source such as glucose or organic acids. When grown in glucose and aromatic hydrocarbons medium, it was observed that the genes for glucose transport and metabolism were downregulated, aromatic hydrocarbons were utilized in the first log phase, and glucose was utilized in the second log phase (Basu et al., 2006; Choudhary et al., 2017). On the other hand, the presence of organic acids did not affect the expression of aromatic hydrocarbon metabolism, so this bacterium is expected to be a candidate strain for biodegradation studies (Phale et al., 2020).
       It is well known that hydrocarbon biotransformation can cause oxidative stress and upregulation of antioxidant enzymes in microorganisms. Inefficient naphthalene biodegradation both in stationary phase cells and in the presence of toxic compounds leads to the formation of reactive oxygen species (ROS) (Kang et al. 2006). Since naphthalene-degrading enzymes contain iron-sulfur clusters, under oxidative stress, the iron in heme and iron-sulfur proteins will be oxidized, leading to protein inactivation. Ferredoxin-NADP+ reductase (Fpr), together with superoxide dismutase (SOD), mediates the reversible redox reaction between NADP+/NADPH and two molecules of ferredoxin or flavodoxin, thereby scavenging ROS and restoring the iron-sulfur center under oxidative stress (Li et al. 2006). It has been reported that both Fpr and SodA (SOD) in Pseudomonas can be induced by oxidative stress, and increased SOD and catalase activities were observed in four Pseudomonas strains (O1, W1, As1, and G1) during growth under naphthalene-added conditions (Kang et al., 2006). Studies have shown that the addition of antioxidants such as ascorbic acid or ferrous iron (Fe2+) can increase the growth rate of naphthalene. When Rhodococcus erythropolis grew in naphthalene medium, the transcription of oxidative stress-related cytochrome P450 genes including sodA (Fe/Mn superoxide dismutase), sodC (Cu/Zn superoxide dismutase), and recA was increased (Sazykin et al., 2019). Comparative quantitative proteomic analysis of Pseudomonas cells cultured in naphthalene showed that upregulation of various proteins associated with the oxidative stress response is a stress coping strategy (Herbst et al., 2013).
       Microorganisms have been reported to produce biosurfactants under the action of hydrophobic carbon sources. These surfactants are amphiphilic surface active compounds that can form aggregates at oil-water or air-water interfaces. This promotes pseudo-solubilization and facilitates the adsorption of aromatic hydrocarbons, resulting in efficient biodegradation (Rahman et al., 2002). Due to these properties, biosurfactants are widely used in various industries. The addition of chemical surfactants or biosurfactants to bacterial cultures can enhance the efficiency and rate of hydrocarbon degradation. Among the biosurfactants, rhamnolipids produced by Pseudomonas aeruginosa have been extensively studied and characterized (Hisatsuka et al., 1971; Rahman et al., 2002). Additionally, other types of biosurfactants include lipopeptides (mucins from Pseudomonas fluorescens), emulsifier 378 (from Pseudomonas fluorescens) (Rosenberg and Ron, 1999), trehalose disaccharide lipids from Rhodococcus (Ramdahl, 1985), lichenin from Bacillus (Saraswathy and Hallberg, 2002), and surfactant from Bacillus subtilis (Siegmund and Wagner, 1991) and Bacillus amyloliquefaciens (Zhi et al., 2017). These potent surfactants have been shown to reduce the surface tension from 72 dynes/cm to less than 30 dynes/cm, allowing for better hydrocarbon absorption. It has been reported that Pseudomonas, Bacillus, Rhodococcus, Burkholderia and other bacterial species can produce various rhamnolipid and glycolipid-based biosurfactants when grown in naphthalene and methylnaphthalene media (Kanga et al., 1997; Puntus et al., 2005). Pseudomonas maltophilia CSV89 can produce extracellular biosurfactant Biosur-Pm when grown on aromatic compounds such as naphthoic acid (Phale et al., 1995). The kinetics of Biosur-Pm formation showed that its synthesis is a growth- and pH-dependent process. It was found that the amount of Biosur-Pm produced by cells at neutral pH was higher than that at pH 8.5. Cells grown at pH 8.5 were more hydrophobic and had higher affinity for aromatic and aliphatic compounds than cells grown at pH 7.0. In Rhodococcus spp. N6, higher carbon to nitrogen (C:N) ratio and iron limitation are optimal conditions for the production of extracellular biosurfactants (Mutalik et al., 2008). Attempts have been made to improve the biosynthesis of biosurfactants (surfactins) by optimizing strains and fermentation. However, the titer of surfactant in the culture medium is low (1.0 g/L), which poses a challenge for large-scale production (Jiao et al., 2017; Wu et al., 2019). Therefore, genetic engineering methods have been used to improve its biosynthesis. However, its engineering modification is difficult due to the large size of the operon (∼25 kb) and complex biosynthetic regulation of the quorum sensing system (Jiao et al., 2017; Wu et al., 2019). A number of genetic engineering modifications have been carried out in Bacillus bacteria, mainly aimed at increasing surfactin production by replacing the promoter (srfA operon), overexpressing the surfactin export protein YerP and the regulatory factors ComX and PhrC (Jiao et al., 2017). However, these genetic engineering methods have only achieved one or a few genetic modifications and have not yet reached commercial production. Therefore, further study of knowledge-based optimization methods is necessary.
       PAH biodegradation studies are mainly conducted under standard laboratory conditions. However, at contaminated sites or in contaminated environments, many abiotic and biotic factors (temperature, pH, oxygen, nutrient availability, substrate bioavailability, other xenobiotics, end-product inhibition, etc.) have been shown to alter and influence the degradative capacity of microorganisms.
       Temperature has a significant effect on PAH biodegradation. As temperature increases, dissolved oxygen concentration decreases, which affects the metabolism of aerobic microorganisms, since they require molecular oxygen as one of the substrates for oxygenases that carry out hydroxylation or ring cleavage reactions. It is often noted that elevated temperature converts the parent PAHs into more toxic compounds, thereby inhibiting biodegradation (Muller et al., 1998).
       It has been noted that many PAH contaminated sites have extreme pH values, such as acid mine drainage contaminated sites (pH 1–4) and natural gas/coal gasification sites contaminated with alkaline leachate (pH 8–12). These conditions can seriously affect the biodegradation process. Therefore, before using microorganisms for bioremediation, it is recommended to adjust the pH by adding suitable chemicals (with moderate to very low oxidation-reduction potential) such as ammonium sulfate or ammonium nitrate for alkaline soils or liming with calcium carbonate or magnesium carbonate for acidic sites (Bowlen et al. 1995; Gupta and Sar 2020).
       Oxygen supply to the affected area is the rate limiting factor for PAH biodegradation. Due to the redox conditions of the environment, in situ bioremediation processes usually require oxygen introduction from external sources (tilling, air sparging, and chemical addition) (Pardieck et al., 1992). Odenkranz et al. (1996) demonstrated that the addition of magnesium peroxide (an oxygen releasing compound) to a contaminated aquifer could effectively bioremediate BTEX compounds. Another study investigated the in situ degradation of phenol and BTEX in a contaminated aquifer by injecting sodium nitrate and constructing extraction wells to achieve effective bioremediation (Bewley and Webb, 2001).