Microcystins (MCs) are the most commonly found cyanotoxins produced by several genera of Cyanobacteria and are the most studied. MCs are cyclic heptapeptides, comprised of cyclo-(D-Ala¹–X²–D-MeAsp³–Z⁴–Adda⁵–D-Glu⁶–Mdha⁷) with approximately 250 identified variants (Spoof and Catherine, 2016; Yang et al., 2020). The X and Z represent variable L-amino acids (microcystins referred as MC-XZ), MeAsp is erythro-β-methylaspartic acid, Adda is (2S, 3S, 8S, 9S) 3-amino-9-methoxy-2,6,8-trimethyl-10-phenyl-deca-4,6-dienoic acid (β-amino acid), and Mdha is N-methyldehydroalanine (Krishnamurthy et al., 1989; Figure 8). Some examples of MC-XZ variants include MC-LR, MC-RR, MC-YR, MC-WR, MC-LY, MC-LW, MC-LF, MC-LA etc., where MC-LR is one of the most widely distributed and highly toxic variant.

MC-LR degradation has been detailed for Sphingopyxis sp. YF1. The cyclic MC-LR is first linearized by cleavage of Adda-Arg peptide bond catalyzed by microcystinase which is further acted upon by linearized microcystinase cleaving the Ala-Leu peptide bond forming a tetrapeptide containing Adda. This tetrapeptide is cleaved at Adda-Glu peptide bond by tetrapeptidase forming Adda. Adda is metabolized to form aromatic aliphatic hydrocarbon (C₂₀H₂₆O₄) by the action of aminotransferase (Figure 8). The aromatic aliphatic hydrocarbon gets converted to phenylacetic acid by the microbial β-oxidation enzymes (fatty acid-CoA ligase, acyl-CoA dehydrogenase, enoyl-CoA hydratase, 3-hydroxyacyl-CoA dehydrogenase, and thiolase) probably in four cycles of β-oxidation releasing acetyl-CoA/propanoyl-CoA in each cycle. Potential intermediates formed during β-oxidation cycles were identified as 7-methoxy-4,6-dimethyl-8-phenyloca-2,4-dienoic acid and 2-methyl-3-methoxy-4-phenylbutyric acid. Such β-oxidation of aromatic aliphatic hydrocarbons has also been reported for alkylbenzenes (Sariaslani et al., 1974; Awe et al., 2008; Nhi-Cong et al., 2010; Figure 8). Further, phenylacetic acid is activated by ligating coenzyme-A catalyzed by fatty acid-CoA ligase, a phenylacetate-CoA ligase like enzyme (PAAase), in Sphingopyxis sp. YF1 and is proposed to be degraded via phenylacetyl-CoA route. The phenylacetyl-CoA is degraded to acetyl-CoA by the action of enzymes phenylacetyl-CoA epoxidase (paaABCDE), 2-(1,2-epoxy-1,2-dihydrophenyl) acetyl-CoA isomerase (paaG), oxepin-CoA hydrolase (paaZ), 3-oxoadipyl-CoA thiolase (paaI) (Figure 8).

The genes involved in MC-LR degradation in strain YF1 include mlrBDAC cluster which converts MC-LR to Adda followed by aminotransferase and microbial β-oxidation encoding genes leading to formation of phenylacetate. Further, the paaI-paaGZ-paaABCDE gene clusters encodes enzymes for degradation of phenylacetate to acetyl-CoA. The mlr and paa clusters along with genes encoding aminotransferase and β-oxidation enzymes were located in proximity suggesting their involvement in MC-LR degradation. The mlr cluster has been observed in other MC-LR degrading microbes such as Sphingosicicella microcystinivorans B-9 and Novosphingobium sp. THN1 (Jin et al., 2018; Wang J. et al., 2019; Yang et al., 2020).

Nodularin is a cyclic pentapeptide comprising of D-MeAsp¹–L-Arg²–Adda³–D-Glu⁴–Mdhb⁵, where 1^st–4^th amino acids are similar to 3^rd–6^th amino acids of MC-LR and the 5^th Mdhb is N-methyldehydrobutyrine. Microcystin degrading bacteria harbouring mlrBDAC cluster have been observed to degrade nodularin, which is a pentapeptide possessing cleavage sites similar to MC-LR (Figure 8). As observed in Sphingopyxis sp. m6, during nodularin degradation mlr cluster was upregulated and products such as linearized nodularin and Adda were detected. This suggests nodularin degradation share similar enzymes/enzymatic steps (Yang et al., 2020; Yuan et al., 2021; Wei et al., 2023).

Di (2-ethylhexyl) phthalate (DEHP) is the most extensively used plasticizer and is a phthalate ester composed of a phthalate backbone with two 2-ethylhexyl groups attached. It is toxic, estrogenic and a potent endocrine disrupting environmental pollutant. Several bacterial genera capable of degrading DEHP have been isolated such as Gordonia, Rhodococcus, Mycrobacterium, Pseudomonas, Cupravidus, Burkholderia, Achromobacter, Agromyces, Microbacterium, Acinetobacter, Bacillus, etc (Zhao et al., 2016; Xu et al., 2017; Zhang et al., 2018; Fan et al., 2018; Li et al., 2019; Wright et al., 2020; Wang et al., 2021; Chen et al., 2021a and b; Kamaraj et al., 2022; Sun et al., 2024). The initial degradation of DEHP typically occurs through two main routes: de-esterification leading to the formation of mono-alkyl esters like mono-(2-ethyhexyl) phthalate (MEHP), or stepwise beta-oxidation of alkyl side chains resulting in DBP (Figure 9). In most of bacterial strains, DEHP is hydrolysed into MEHP by esterases, which is further converted to phthalic acid, either directly or via mono-butyl phthalate (MBP) (Ren et al., 2016; Xu et al., 2017; Nahurira et al., 2017; Fan et al., 2018; Li et al., 2019; Lamraoui et al., 2020; Zhang H. et al., 2020; Wang et al., 2021; Kamaraj et al., 2022; Hsu et al., 2023; Kou et al., 2023; Bhattacharyya et al., 2023; Dhar et al., 2023). In other strains, the alkyl side chain of DEHP is first oxidised to DBP, which is then hydrolysed to yield phthalic acid, either directly (Zhang et al., 2021; Chen et al., 2021a; Chen et al., 2022) or through intermediates like diethyl phthalate (DEP), mono-methyl phthalate (MMP) or butyl methyl phthalate (BMP), MBP (Chen et al., 2021b; Zhang et al., 2021; Figure 9).

Some strains employ both de-esterification (via MEHP) and alkyl side chain oxidation (via DBP) routes to degrade DEHP to phthalic acid (Zhao et al., 2018; Chen et al., 2021b; Chang et al., 2022). A few strains such as Rhodococcus pyridininvorans DNHP-2 exhibit alternate pathways wherein DEHP undergoes conversion to 2-ethyl hexyl benzoic acid (2EHBA), which is further converted into benzoic acid (Wang et al., 2022). In Gordonia sp. LFF, DEHP is metabolized to phthalic acid via ethylhexyl hexyl phthalate (EHHP), butyl-(2-ethylhexyl) phthalate (BEHP), MEHP, mono-hexyl phthalate (MHP), and MBP (Wang et al., 2019b). In Microbacterium sp. DEHP1 and Mycolibacterium phocacium RL-HYO1, DEHP is converted to phthalic acid through intermediates di (2-ethylbutyl) phthalate (DEBP), di-n-hexyl phthalate (DnHP), DBP, and diethyl phthalate (DEP) (Ren et al., 2021; Sun et al., 2024; Figure 9).

Further, the resulting phthalic acid is converted to protocatechuate (PCA) either via 3,4-dihydroxyphthalate (34DHP) (Fan et al., 2018; Zhao et al., 2018; Chen et al., 2021a; Chen et al., 2021b; Bhattacharyya et al., 2023) or 4,5-dihydroxyphthalate (45DHP) (Xu et al., 2017). In few strains phthalic acid is converted to salicylate and then to gentisate or catechol (Chen et al., 2007; Ren et al., 2021). While in some bacteria, phthalic acid is converted to benzoic acid and then funneled to catechol (Chen et al., 2021a; Wang et al., 2022; Sun et al., 2024). Common intermediates like PCA, gentisate and catechol are then ring cleaved by dioxygenases and subsequently funneled into the TCA cycle (Figure 9).

Dibutyl phthalate (DBP) is a plasticizer that is extensively used in the production of PVC products, such as flexible plastics, vinyl flooring, and medical devices. It exhibits severe endocrine-disrupting properties as well as liver and respiratory toxicity. Several bacterial genera such as Bacillus, Acinetobacter, Pseudomonas, Mycobacterium, Halomonas, Cupravidus, Arthrobacter, Microbacterium, among others, have been reported to degrade DBP (Feng et al., 2018; Wright et al., 2020; Feng et al., 2021; Chen et al., 2021b; Nandi et al., 2021; Li et al., 2022; Sun et al., 2024). The degradation of DBP typically begins with its hydrolysis to mono-butyl phthalate (MBP) by esterases, either directly (Kumar and Maitra, 2016; Feng et al., 2018; Xu et al., 2022; Shariati et al., 2022; Fan et al., 2023; Sun et al., 2024) or via intermediates like butyl ethyl phthalate (BEP) and butyl methyl phthalate (BMP) (Feng et al., 2021; Mondal et al., 2024). Alternatively, DBP is converted to phthalic acid through intermediates such as diethyl phthalate (DEP) or dimethyl phthalate (DMP) (Sun et al., 2019; Mondal et al., 2024; Figure 9). Alternatively, in some strains, such as Pseudomonas aeruginosa PS1 and Halomonas sp. ATBC28, DBP is converted to butyl benzoate, which is further metabolized to benzoic acid (Wright et al., 2020; Du et al., 2024).

Furthermore, the resulting phthalic acid is converted to PCA via intermediates like 34DHP (Feng et al., 2018; Liu et al., 2020; Wright et al., 2020; Nandi et al., 2021; Chen et al., 2021b) or 45DHP (Feng et al., 2021; Du et al., 2024). In certain bacterial strains, phthalic acid is converted to benzoic acid, which undergoes decarboxylation to yield catechol. For instance, Glutamibacter sp 0426, Enterobacter DNB, and Arthrobacter ZJUTW convert phthalic acid to PCA via benzoic acid (Sun et al., 2019; Liu et al., 2020; Ren et al., 2023b; Figure 9). Similarly, in Pseudomonas YJB6, phthalic acid is converted to PCA via benzoic acid, 45DHP, and catechol (Feng et al., 2021). Paenarthrobacter ureafaciens PB10 converts phthalic acid into gentisate via 4-hydroxyphthalic acid (4HP) (Shariati et al., 2022). Common intermediates like PCA, gentisate and catechol are then subjected to ring cleavage, facilitating their entry into the tricarboxylic acid (TCA) cycle (Figure 9).

Benzyl butyl phthalate is a plasticizer that is composed of a phthalate backbone with a benzyl group and a butyl group attached to it. It is widely used in synthesis of various industrial and consumer products like PVC pipes, rubber, adhesives, cosmetics and has been reported to demonstrate reproductive and developmental toxicity, endocrine disruption, etc. Various bacterial genera, including Bacillus, Acinetobacter, Arthrobacter, Gordonia, and others, have been reported with the ability to degrade benzyl butyl phthalate (BBP) (Chatterjee and Dutta, 2003; Zhang et al., 2018; Nandi et al., 2021; Kaur et al., 2021; Fan et al., 2023). These organisms typically employ two primary pathways for BBP degradation. In the first pathway, esterases hydrolyze the alkyl side chain of BBP, yielding mono-benzyl phthalate (MBeP), which is further metabolized to phthalic acid and benzyl alcohol (Chatterjee and Dutta, 2003; Zhang et al., 2018; Kaur et al., 2021; Fan et al., 2023; Figure 9). Alternatively, in the second pathway, the aromatic side chain of BBP undergoes hydrolysis, resulting in the formation of benzyl alcohol and mono-butyl phthalate (MBP), which is then converted to phthalic acid (Chatterjee and Dutta, 2003; Zhang et al., 2018; Nandi et al., 2021; Kaur et al., 2021; Fan et al., 2023; Figure 9). Both pathways yield benzyl alcohol, which is subsequently metabolized to catechol via benzoic acid while the resulting phthalic acid is metabolized to PCA either via benzoic acid (Zhang et al., 2018) or via 34DHP and 3,4-dihydroxybenzoic acid (Kaur et al., 2021; Figure 9).

Di-n-octyl phthalate (DnOP) is a type of phthalate ester commonly used as a plasticizer in various industrial applications and is composed of two octyl groups attached to a phthalate backbone. It is known to be a potent endocrine disruptor, carcinogen and immunotoxin. Several bacterial genera capable of degrading DnOP have been identified, including Arthrobacter, Rhodococcus, Gordonia, Burkholderia, Bacillus, among others (Wu et al., 2010; Sarkar et al., 2013; Zhang et al., 2017b; Zhang et al., 2018; Gani and Kazmi, 2018; Dhar et al., 2023). In most of bacterial strains, DnOP degradation begins with the hydrolysis by diesterase to yield mono-n-octyl phthalate (MnOP) which is subsequently converted to phthalic acid (Sarkar et al., 2013; Zhang et al., 2017b; Zhang et al., 2018; Dhar et al., 2023; Figure 9). Alternatively, in a co-culture of Gordonia sp. JDC-2 and Arthrobacter sp. JDC-32, DnOP degradation occurs through sequential β-oxidation of the alkyl side chain of DnOP, leading to the formation of intermediates such as hydroxyl octyl phthalate (HOP), bis(2-oxoheptyl) phthalate (BOP), DBP and DEP. DEP is then hydrolyzed to yield MMP, which is subsequently converted to phthalic acid (Wu et al., 2010; Figure 9). Similarly, in the halotolerant consortium LF, DnOP degradation was initiated by β-oxidation of the alkyl side chain to yield intermediates such as HOP, BOP or dihexyl phthalate (DiHP), bis(2-hydroxypropyl) phthalate (BHP), DBP, and MBP (Wang et al., 2020). The resulting phthalic acid was converted to PCA either via dihydroxyphthalates (Zhang et al., 2018; Dhar et al., 2023) or benzoic acid (Zhang et al., 2018; Figure 9).

Neonicotinoids are a recent class of pesticides consisting of thiacloprid, acetamiprid, imadacloprid, clothianidin, etc., used for crop protection, horticulture, and flea control. Imidacloprid is the most commonly used insecticide of the neonicotinoid group and is recognized as a contaminant of emerging concern (Selvam and Srinivasan, 2019; Petkovic Didovic et al., 2022). Low bioavailability of imidacloprid results in slow rate of natural attenuation and leads to longer half-life (∼997 days) in soil. Extensive use of imidacloprid exerts adverse effects on non-target species like fish, bees, earthworm, mice, human, etc. (Phugare et al., 2013; Pang et al., 2020).

Various bacteria including Pseudomonas, Bacillus, Klebsiella, Mycobacterium, etc., have been reported to degrade imidacloprid with various possible pathways. Among the reported routes, oxidation and nitro-reduction are two major microbial biodegradation pathways of imidacloprid (Pandey et al., 2009; Phugare et al., 2013; Pang et al., 2020; Zhang X. et al., 2023). 6-Chloronicotinic acid (6-CNA), olefinic cyclic nitroguanidine, cyclic urea, cyclic guanidine, nitroso, and nitro derivatives are major metabolites of imidacloprid nitro-reduction detected in soil and water samples. Under microaerophilic conditions, an aldehyde oxidase converts the ‘magic nitro’ group of imidacloprid to a nitrosoguanidine metabolite. The imidacloprid and/or formed product i.e., nitrosoguanidine is degraded through a more toxic nitroguanidine intermediate which is further converted into non-toxic urea metabolites (Pandey et al., 2009; Figure 10). In some microorganisms, imidacloprid is cleaved to 6-CNA by the formation of nitrosoguanidine and oxidative cleavage of guanidine residue (Phugare et al., 2013). Alternatively, imidacloprid is converted to 6-CNA via formation of 5-hydroxy and olefin metabolites by subsequent hydroxylation and dehydrogenation. 6-CNA is eventually converted to CO₂ via 6-hydroxynicotinic acid (Sharma et al., 2014; Figure 10).

Chlorpyrifos, [O,O-diethyl O-(3,5,6-trichloro-2-pyridinyl)-phosphorothioate], is a broad-spectrum, chlorinated organophosphate insecticide, acaricide and miticide used to control foliage- and soil-borne insect pests on a variety of food and feed crops (Lara-Moreno et al., 2022; Bosu et al., 2024). The major health issues caused by chlorpyrifos include respiratory, immunological, reproductive, and neurological disorders in humans (Anwar et al., 2009). Although potential health risks have led to the ban of chlorpyrifos in many countries, it has been approved for limited use in densely populated countries like India, China and Bangladesh (Lara-Moreno et al., 2022). In Mexico, for example, extensive use of chlorpyrifos from 2012 to 2020 resulted in contamination of waterbodies (estuaries, drains and artesian wells) with an average concentration of 4,614 ng L⁻¹ of chlorpyrifos (Ruiz-Arias et al., 2023). Chlorpyrifos has an average half-life of around 60–120 days in the soil, depending upon climate and soil stability (Anwar et al., 2009; Bosu et al., 2024). Various microorganisms belonging to the genera Arthrobacter, Enterobacter, Xanthomonas, Streptomyces, Stenotrophomonas, Sphingomonas, Bacillus, Synechocystis, Pseudomonas, Actinobacteria, and Klebsiella have been identified as potential chlorpyrifos degraders (Singh, 2009; Ambreen and Yasmin, 2021).

Various microorganisms are known to produce metal-dependant enzymes (hydrolases) such as organophosphorus hydrolase, phosphotriesterase (PTE), methyl parathion hydrolase and organophosphorus acid anhydrolase (OPAA) involved in chlorpyrifos bioremediation (John and Shaike, 2015; Bosu et al., 2024). Organophosphorus hydrolase effectively cleaves P–O bond in the phosphotriesters, and P–S linkage in the phosphothiolesters, yielding two major metabolites, 3,5,6-trichloro-2-pyridinol (TCP) and diethylphosphate (DETP) from chlorpyrifos. Other minor metabolites such as desethyl chlorpyrifos, chlorpyrifos oxon, desethyl chlorpyrifos oxon, and 3,5,6-trichloro-2-methoxypyridine (TMP) are also produced. Chlorpyrifos oxon, the oxidized form of chlorpyrifos, is further hydrolyzed either enzymatically or spontaneously to form diethylphosphate and TCP. The TCP can be further degraded to TMP and CO₂ (Racke, 1993; John and Shaike, 2015; Figure 10).

Carbendazim (methyl N-(1H-benzimidazol-2-yl) carbamate) is a systemic broad-spectrum fungicide, which is also formed as a degradation product of thiophanate-methyl and benomyl fungicides (Mazellier et al., 2003; Fang et al., 2010). Carbendazim is used worldwide as pre- and post-harvest treatment to control the Ascomycetes, Fungi imperfecti and Basidiomycetes fungal diseases on various vegetables, fruits and several other plants. Carbendazim was found to be toxic to various animals and could induce reproductive, developmental, endocrine and haematological toxicity (Rama et al., 2014; Zhou et al., 2023). Many microorganisms, predominantly bacteria such as Rhodococcus, Nocardioides, Mycobacterium, Pseudomonas, etc., have been reported to metabolize carbendazim (Singh et al., 2016; Zhou et al., 2023). Among reported microbes, few bacterial strains have been found to be efficient degraders of carbendazim. For example, Rhodococcus sp. D-1 isolated from contaminated farmland in China, could degrade 98.20% of 200 ppm carbendazim within 5 days (Bai et al., 2017). In most organisms, carbendazim degradation is initiated by its hydrolysis to 2-aminobenzimidazole (2-AB) and further, 2-hydroxybenzimidazole (2-HB) (Wang et al., 2010; Arya et al., 2015; Figure 10). Subsequently, 2-HB is converted to catechol via the formation of benzene-1,2-diamine and further channelled into TCA cycle (Singh et al., 2016; Figure 10).

Chloroacetanilide herbicides such as alachlor, metolachlor, and acetochlor are primary herbicides, and more than 50 million kg has been used annually in the United States (Gan et al., 2002). These herbicides are highly soluble in water and persist in soil, with residues or metabolites being detected in surface and ground water (Potter and Carpenter, 1995; Tian et al., 2021). For example, alachlor was detected in groundwaters in the United States at concentration 0.1–16.6 μg L⁻¹, exceeding the U.S. Environmental Protection Agency (US-EPA) maximum contaminant level criteria of 2 μg L⁻¹ (WHO, 2017).

Alachlor [2-chloro-N-(2,6-diethylphenyl)-N-(methoxymethyl]acetamide] is one of the majorly used chloroacetanilide applied as a selective pre- and post-emergent herbicide to control weeds in soybeans, peanuts, and corn crops. C-dealkylation of other chloroacetanilide herbicides like butachlor leads to formation of alachlor. Alachlor has been categorised as a human carcinogen and has been reported to mimic 17β-estradiol, thereby acting as an endocrine-disruptor (Lee and Kim, 2022). Microbes including various genera of bacteria like Paracoccus, Rhodococcus, Pseudomonas, Acinetobacter, Streptomyces, etc., and few fungi like Paecilomyces have been reported to degrade alachlor (Słaba et al., 2013; Lee and Kim, 2022; Chen et al., 2023). Alachlor is converted to 2-chloro-N-(2,6-diethylphenyl) acetamide (CDEPA) by N-dealkylation (Zhang et al., 2011; Figure 10). Various hydrolases/reductases have been reported to be involved in N-dealkylation of chloroacetanilides. For example, enzyme ChlH from Rhodococcus sp. B1 and enzymes CndB1, CndB2, and CndC1 from Sphingomonas sp. DC-6 have been reported to catalyse the N-dealkylation of alachlor as well as other chloroacetamide like acetochlor, butachlor, and pertilachlor (Chen et al., 2023; Figure 10). Subsequently, CDEPA is transformed to 2,6-diethylanilline (DEA) with the help of enzyme amidase (CmeH) or amide hydrolase (DamH). DEA is further converted to aniline which is then acted upon by aniline dioxygenase to form catechol. The formed catechol is then oxidized through an ortho-cleavage pathway to cis, cis-muconic acid and funnelled into TCA cycle (Zhang et al., 2011; Kim et al., 2013; Gao et al., 2015; Figure 10).

Glyphosate is a low-molecular-weight phosphonate (non-specific organophosphate herbicide) with high aqueous solubility and mobility, which leads to rapid leaching of this compound into soil, causing contamination of water bodies. Glyphosate has been reported to cause toxicity to bacteria as well as multicellular organisms like non-target crop plants, crustaceans, molluscs and chordates including humans (cytotoxicity and genotoxicity) (Zhan et al., 2018; Singh et al., 2020). Various strains of Achromobacter, Agrobacterium, Pseudomonas, Ochrobactrum, etc., have been previously isolated from contaminated sites which can utilize glyphosate as growth substrate (Zhao et al., 2015; Zhan et al., 2018; Feng D. et al., 2020). The primary degradation products of glyphosate include aminomethylphosphonic acid (AMPA) and sarcosine, which are reported to be more toxic than the parent compound (Zhan et al., 2018; Lozano and Pizarro, 2024; Figure 10). C-P lyase removes phosphate group from glyphosate yielding sarcosine, which is cleaved by sarcosine oxidase (encoded by 7 sox genes) into glycine and formaldehyde. Both these intermediates are funnelled into microbial metabolism and biosynthetic pathways (González-Valenzuela and Dussán, 2018; Diaz-Soto et al., 2024; Figure 10). In many microorganisms, glyphosate is converted to AMPA and glyoxylate by the action of glyphosate oxidoreductase. Further, glyoxylate is metabolized to TCA cycle, whereas AMPA is either acted upon by C-P lyase to produce methylamine or by an aminotransferase to form formylphosphonate. Both the intermediates i.e., methylamine and formylphosphonate are cleaved to formaldehyde, which is used by microbes for biosynthesis (Sviridov et al., 2015; Zhao et al., 2015; Singh et al., 2019; Figure 10).

The application of various omics approaches to CEC degradation can provide possible systemic-level insights into the metabolic pathways and associated regulatory mechanisms. Genomics aids in identifying key genes encoding degradative enzymes and/or the evolutionary trajectory. For example, in strain Pseudomonas sp. C5pp, the draft genome analysis revealed the presence of three gene clusters on a single contig (Supercontig-A) involved in complete Carbaryl degradation. The genome analysis further suggested acquisition of genes by horizontal gene transfer events (Trivedi et al., 2016). Genomics in conjunction with transcriptomics and proteomics aids in identifying up/downregulation of genes/proteins under target conditions. For example, in P. bharatica CSV86^T, the transcription analysis showed the induction of target genes involved in benzoate (benE, benK) and glucose (gbp, oprB, glcG) utilisation (Choudhary et al., 2017). Further proteomic analysis aided in identification of Gbp and OprB as glucose binding protein and carbohydrate specific porin, respectively which are induced when the culture is grown on glucose (Basu et al., 2007; Shrivastava et al., 2011). In Methylorubrum sp. ZY-1, the integrated transcriptomic and metabolomic analyses aided in revealing degradation of pentachlorodiphenyl (PCB 118) and underlying molecular mechanisms (Wu et al., 2024). Table 2 provides a comprehensive review of the application of omics in CEC biodegradation for pharmaceuticals, cyanotoxins, plasticizers and pesticides, which are also described further.

Comparative transcriptomic analyses of four penicillin degrading strains revealed upregulation of genes encoding beta-lactamase (bla), penicillin amidase (put) and phenylacetic acid degradation enzymes (paa) in penicillin grown cells (as compared to alternative carbon source grown cells). In Pseudomonas sp. strain ABC07, the put operon, encoding four open reading frames: a beta-lactamase, a major facilitator family importer and two amidases (put1 and put2) was found to be upregulated in presence of penicillin and benzylpenicilloic acid, but not phenylacetic acid. Whereas, the paa operon was responsive to all three intermediates (penicillin, benzylpenicilloic acid and phenylacetic acid). Therefore, these analyses indicated the metabolic architecture of penicillin degradation in strain ABC07 (Crofts et al., 2018). A combination of metagenomic and cultivation-based techniques identified Arthrobacter and Pimelobacter as the dominant members of a sulphadiazine-degrading consortia as well as their individual roles in degradation. While Arthrobacter sp. D2 converted sulphadiazine to 2-aminopyrimidine, Pimelobacter sp. LG209 mineralised this intermediate (Deng et al., 2018). Exposure of Microbacterium sp. C448 to therapeutic and sub-therapeutic doses of sulphamethazine was assessed using transcriptomic and proteomic analyses. The transcript and protein levels of the degradation enzymes monooxygenase SadA and flavin reductase SadC exhibited increase in presence of sulphamethazine. Further, the upregulation of the enzyme RidA (reactive intermediate deaminase A) suggested its potential role in deamination of 2-aminophenol. Additionally, the putative sulphate exporter family protein showed upregulation in presence of sulphamethazine (Paris et al., 2023). An integrated multi-omics approach revealed the chloramphenicol biotransformation pathway, genes, proteins/enzymes and community structure/interactions of the activated sludge enriched consortium CL. The metagenomic analysis revealed Sphingomonas, Caballeronia, Cupriavidus and Pigmentiphaga as the major players in chloramphenicol degradation. Further, metatranscriptomic analysis revealed upregulation of specific detoxification and metabolic pathway genes such as capO, which encodes a glucose-methanol-choline oxidoreductase responsible for oxidation of C₃-OH group of chloramphenicol. The proteomic analysis validated the metatranscriptomic data and the functionality of the identified enzymes such as CapO, nitroreductase and chloramphenicol acetyltransferase (Zhang et al., 2022).

The genomic and proteomic analysis of sulphamethoxazole-degrading Pseudomonas silesiensis F6a revealed six key degradation genes, deoC (2-deoxyribose 5-phosphate aldolase), narI (nitrate reductase), luxS (S-ribosyl homocysteine lyase), nuoH (NADH quinone oxidoreductase), gene 0655 (F420 dependent oxidoreductase) and gene 4,650 (amidohydrolase) involved in C-S bond cleavage, S-N bond hydrolysis and isoxazole ring-cleavage (Liu et al., 2022).

Metabolomic analyses of Sphingobacterium mizutaii S121 revealed the products of tetracycline biodegradation by the strain and the stress response mechanisms involved. Based on the analyses, two biodegradation pathways involving demethylation and one hydrolysis pathway were proposed. The levels of indole, glutamic acid and FAD, involved in regulating the activity of efflux proteins and degradation enzymes, were upregulated. Further, intracellular levels of nucleotides and amino acids were significantly increased to repair DNA/RNA and protein in response to tetracycline stress. The levels of antioxidants such as taurine and protoporphyrin IX also increased in response to the generation of reactive oxygen species due to enhanced aerobic metabolism. Under tetracycline stress, strain S121 required increased nutrients from the extracellular environment, due to which the levels of the metabolite N-palmitoyl sphingomyelin and phosphoethanolamine decreased significantly, enhancing membrane fluidity (Tan H. et al., 2022).

A metagenomic analysis of sewage samples incubated with testosterone indicated the genera Comamonas and Pseudomonas to be major players in degradation. Further, the meta-cleavage dioxygenase gene tesB was identified and exhibited a significant increase after 48 h of incubation (Chen et al., 2016). A combination of metagenomic analyses and comparative genomics revealed Actinobacteria and Proteobacteria as major players in oestrogen biodegradation distributed in aquatic ecosystems. Further, aedJ and edcC were identified as biomarkers for oestrogen degradation in Actinobacteria and Proteobacteria, respectively, with potential application for environmental detection (Hsiao et al., 2023).

Quantitative proteomics of Patulibacter sp. I11 in absence or presence of ibuprofen revealed likely proteins involved in degradation. In presence of ibuprofen, various proteins involved in aromatic degradation such as enoyl-CoA hydratase/isomerase, acyl-CoA synthetase, Rieske (2Fe-2S) domain containing were upregulated. Additionally, other proteins such as ABC transporter (probably involved in ibuprofen uptake), putative lyase, stress response protein and AMP-forming synthetase were also upregulated (Almeida et al., 2013). The stress response of Serratia nematodiphila sp. MB307 to the presence of ibuprofen was investigated using differential proteomics. Thirteen proteins were upregulated and 29 proteins were downregulated in response to ibuprofen stress. Among the upregulated proteins, Fe-S cluster scaffold-like protein IscU, autoinducer-2 modifying protein LsrG and peptidylprolyl isomerase have been implicated for their role in stress tolerance. Overall, the analyses highlighted the multifaceted stress response of strain MB307, involving a balance between protein synthesis, DNA replication, and energy production (Basharat et al., 2022).

The genomic-proteomic-metabolomic analyses of Paracoccus sp. APAP_BH8 elucidated the genes, enzymes and metabolic pathway of acetaminophen degradation in the strain. The proteome analysis revealed the upregulation of M20 aminoacylase (amidohydrolase), guanine deaminase GuaD, 4-hydroxybenzoate-3-monoxygenase PobA and 4-hydroxyphenyl pyruvate dioxygenase HppD in presence of acetaminophen. Molecular docking studies of these enzymes with their respective substrates validated the functionality of these enzymes. Further, the metabolomic analysis revealed 4-aminophenol, hydroquinone and 3-hydroxy cis, cis-muconate as degradation pathway intermediates (Pandey et al., 2024).

Genomic and comparative transcriptome analyses of Aminobacter sp. Strain NyZ550 revealed upregulation of the genes dmm (dimethylamine monooxygenase), gmas (γ-glutamylmethylamide synthetase), mgs (N-methylglutamate synthase) and mgd (N-methylglutamate dehydrogenase) while growing on metformin, indicating the metabolism of dimethylamine. Further, the serine cycle and formate-tetrahydrofolate catabolic genes also exhibited upregulation. Importantly, the gene encoding agmatinase (putative metformin hydrolase) exhibited upregulation, highlighting its role in degradation (Li et al., 2023).

In Sphingopyxis sp. YF1, genomic analysis revealed the presence of MC degrading cluster mlrBDAC. The MC-LR degradation products such as linearized MC-LR, tetrapeptide, Adda and its degradation intermediates, and phenylacetic acid were detected using UPLC and UPLC-ESI-MS. Further, transcriptomics and qRT-PCR analyses suggested the upregulation of mlr cluster, fatty acid β-oxidation genes and paa cluster during MC-LR degradation. Metabolomics study showed enrichment of metabolites in pantothenate and CoA biosynthesis as well as fatty-acid degradation indicating involvement of fatty acid β-oxidation in MC-LR degradation (Wei et al., 2023). In Sphingopyxis sp. m6, gene specific qPCR suggested involvement of mlr cluster in nodularin degradation and the respective enzymatic steps were identified by metabolite analysis using total ion chromatogram (Yuan et al., 2021).

In Gordonia sp. GONU, genome sequencing aided in identification of key esterases involved in initial hydrolysis DOP and DEHP to phthalate as well as pht gene clusters (responsible for conversion of phthalate to protocatechuate) and pca gene clusters (responsible for conversion of protocatechuate to TCA cycle intermediates). Substrate dependent gene expression profile by qRT-PCR and protein profiling by LC-ESI-MS/MS revealed that esterases EstG5 and EstG3 are specifically expressed to hydrolyse DnOP to PA, whereas EstG2 and EstG3 are specifically expressed to metabolise DEHP to PA via MEHP (Dhar et al., 2023).

Genome sequencing of Microbacterium sp. DEHP1 identified two key esterases-est258 and monooxygenase 0132 as well as three operons (ben, cat, pca) involved in the degradation of DBP and DEHP. Genome mining and metabolite identification by GC-MS suggested that DBP could be hydrolyzed by esterase 2,518 to yield mono-butyl phthalate (MBP) and subsequently phthalic acid (PA) whereas DEHP may be converted to di-n-hexyl phthalate (DnHP) and then DBP by monooxygenase 0132. Further, metabolic profiling using UHPLC-QTOF/MS revealed that under DEHP/DBP stress, strain DEHP1 cells showed increased levels of valine (which induces production of osmoregulatory substances), glycerophospholipids (major component of cell membrane), glutathione/protoanemonin (antioxidants), and proline (key player to preserve cellular glutathione redox status by activating signaling pathway). Notably, levels of organic substances like levan and naringenin 4′-O-alpha-L-rhamnopyranoside decreased in response to DEHP stress (Sun et al., 2024).

A combined genomic and transcriptomic approach identified 66 key genes involved in two different mono-butyl phthalate-catabolism steps in Pseudomonas aeruginosa PS1. In addition to the genes encoding the metabolic pathway enzymes, most differentially expressed genes in Pseudomonas aeruginosa PS1 under DBP stress were those encoding for ABC transporters, two-component systems, biofilm formation, quorum sensing and chemotaxis (Du et al., 2024).

The genome sequencing of DBP degrading Arthrobacter sp. ZJUTW identified the presence of pehA gene encoding a DBP-hydrolyzing esterase and pht gene cluster responsible for converting phthalic acid to protocatechuate located on a plasmid, and pca gene clusters responsible for converting PCA to TCA cycle intermediates located on the chromosome. Additionally, transcriptomic analysis by RNA-seq showed the upregulation of genes encoding an alpha-ketoglutarate transporter (important for cell wall synthesis), chaperones, MFS transporters (important for DBP efflux), flavin-dependent oxidoreductases, and NADPH-dependent FMN reductase genes (Liu et al., 2020).

Proteogenomic and metabolomic analysis of Halomonas sp. ATBC28 and Mycobacterium sp. DBP42 identified key esterases for DBP hydrolysis (cutinase 0019 in strain DBP42, esterase 4,375 in strain ATBC28) and pht-pca/ben-cat gene clusters. Metabolite analysis revealed that DBP degradation involves sequential removal of the ester-bound side chains, producing phthalate and butanol, while DEHP follows a sequential shortening of the side chains. Further, in strain ATBC28, active membrane transporters (TRAP transporters 0264 and 0631) and a membrane-linked OmpA-like protein (3,348) were upregulated, potentially for detoxification and biosurfactant production, respectively (Wright et al., 2020).

Proteomics and metabolomics provided an enhanced understanding of alachlor biodegradation by P. marquandii. Metabolomics (by LC-MS/MS) suggested that presence of alachlor reduced the culture growth and glucose consumption rates and increased the formation of supplementary materials (UDP-glucose/galactose) and ROS scavengers (ascorbate). Proteomic analysis (2-D electrophoresis and MALDI-TOF/TOF) revealed that the presence of alachlor led to upregulation of enzymes related to energy, sugar metabolism and ROS production. Further, overexpression of cyanide hydratase implicated the key role of this enzyme in the alachlor biodegradation pathway (Szewczyk et al., 2015).

The genomic analysis of Burkholderia cenocepacia CEIB S5-2 revealed the presence of key genes involved in glyphosate degradation pathways (sarcosine and AMPA pathway), suggesting that the bacterial strain could use both routes for glyphosate degradation. Genes soxA/B/D/G encoding sarcosine oxidase enzyme involved in sarcosine pathway as well as genes gox encoding glyphosate oxidoreductase, mauE/D encoding methylamine dehydrogenase and other genes encoding aminotransferases, phosphonatase enzymes involved in AMPA pathway were present on the genome (Diaz-Soto et al., 2024). Genomic data analysis of Lysinibacillus sphaericus suggested the presence of sarcosine oxidase gene and qRT-PCR analysis showed upregulation of this gene in presence of glyphosate (González-Valenzuela and Dussán, 2018).

Degradation pathway of chlorpyrifos and glyphosate in Bacillus cereus strains AKAD 3–1 were elucidated by GC-MS based metabolomics. Analysis of the intermediate and the final metabolic products confirmed that no toxic compounds were produced during chlorpyrifos and glyphosate degradation. This indicates that the bacterium harbors the metabolic pathway for detoxification and degradation of chlorpyrifos and glyphosate into non-toxic compounds (Malla et al., 2023).

Genomic analysis of Sphingobacterium sp. InxBP1 indicated the presence of various genes encoding enzymes involved in imidacloprid degradation. For example, nitronate monooxygenase (locus id K7A41_01745), amidohydrolase family enzymes or metal-dependent hydrolases (K7A41_03835, K7A41_07535) having similarity with 6-chloronicotinic acid chlorohydrolase, and FAD dependent monooxygenase (K7A41_12,275) similar to 6-hydroxy nicotinate monooxygenase, were found to be present in the genome, indicating the potential of strain InxBP1 to degrade imidacloprid (Gautam et al., 2023).

Therefore, omics techniques provide crucial data on various factors such as genes/proteins/metabolites involved, microbial community structure/dynamics, gene expression regulation and stress response mechanisms, thereby aiding in rational design of bioremediation and metabolic engineering strategies.

Application of natural isolates for bioremediation of CECs might face limitations such as slow degradation rates, incomplete transformation into toxic by-products, reduced survivability, and presence of simple carbon sources in the environment (Nielsen, 2001; Dvořák et al., 2017; Phale et al., 2020). These limitations can be overcome by directed genetic engineering approaches, called as “metabolic engineering”. These techniques can be used to broaden metabolic diversity, enhance degradation rates, enhance physiological vigour, overcome carbon catabolite repression etc. (Dvorak et al., 2017).

Multiple reports have described metabolic engineering of CEC degradation/transformation pathways in bacteria. For example, plasmid-mediated expression of sulfonamide monooxygenase and flavin reductase rendered E. coli BL21 (DE3) resistant to sulfamethoxazole (Kim D. et al., 2019). E. coli strain W, carrying the phenylacetic acid catabolic genes (paa operon) was engineered for penicillin utilisation by expression of beta-lactamase and penicillin amidase (pga) (Crofts et al., 2018). A consortium of Aminobacter sp. NyZ550 (that converts metformin to guanylurea) and metabolically engineered P. putida PaW340 (expressing guanylurea hydrolase GuuH) was constructed for metformin mineralisation. While strain NyZ550 converted metformin to guanylurea and dimethylamine, strain PaW340 metabolised guanylurea to guanidine, which was used as nitrogen source by strain NyZ550 (Li et al., 2023). Table 3 provides a comprehensive review of the application of metabolic engineering to CEC biodegradation.