Burning question: Rethinking organohalide degradation strategy for bioremediation applications

Abstract

Organohalides are widespread pollutants that pose significant environmental hazards due to their high degree of halogenation and elevated redox potentials, making them resistant to natural attenuation. Traditional bioremediation approaches, primarily relying on bioaugmentation and biostimulation, often fall short of achieving complete detoxification. Furthermore, the emergence of complex halogenated pollutants, such as per‐ and polyfluoroalkyl substances (PFASs), further complicates remediation efforts. Therefore, there is a pressing need to reconsider novel approaches for more efficient remediation of these recalcitrant pollutants. This review proposes novel redox‐potential‐mediated hybrid bioprocesses, tailored to the physicochemical properties of pollutants and their environmental contexts, to achieve complete detoxification of organohalides. The possible scenarios for the proposed bioremediation approaches are further discussed. In anaerobic environments, such as sediment and groundwater, microbial reductive dehalogenation coupled with fermentation and methanogenesis can convert organohalides into carbon dioxide and methane. In environments with anaerobic‐aerobic alternation, such as paddy soil and wetlands, a synergistic process involving reduction and oxidation can facilitate the complete mineralization of highly halogenated organic compounds. Future research should focus on in‐depth exploration of microbial consortia, the application of ecological principles‐guided strategies, and the development of bioinspired‐designed techniques. This paper contributes to the academic discourse by proposing innovative remediation strategies tailored to the complexities of organohalide pollution.

This review specifically focuses on the novel solution of organohalide bioremediation, proposing redox‐potential‐mediated hybrid bioprocesses tailored to the complexities of organohalide pollution.

INTRODUCTION

Organohalides are organic compounds containing one or more halogens (e.g. fluorine/F, chlorine/Cl, bromine/Br), which have been identified from different environments of both anthropogenic and natural sources (Gribble, 2023; Schymanski et al., 2023). The anthropogenic organohalides represent a wide array of chemicals and are among the most environmentally pervasive and long‐standing chemicals (He et al., 2021; Miner et al., 2017; Potapowicz et al., 2020). Particularly, organohalides being relevant to environmental toxicology include organofluorines, organochlorines and organobromines, which form the structural basis for several environmental contaminants of concern, for example, polychlorinated biphenyls/PCBs, organochlorine pesticides, poly and perfluorinated substances/PFASs, polybrominated diphenyl ethers/PBDEs and brominated flame retardants (Birnbaum & Staskal, 2004; Desforges et al., 2018; Hites, 2004; Keswani et al., 2022; Wang et al., 2017). These chemicals are massively produced for industrial and agricultural purposes, including pesticides, paints, cookware, textiles, personal care products and flame retardants (Agarwal et al., 2017). Due to their widespread applications, organohalides can leach into the environment and are detectable in various environmental matrices (e.g. surface water, sediment, soil, air) at levels posing risks to wildlife and human (He et al., 2021). Moreover, the most concerning issue is the relative resistance of organohalides to degradation, resulting in their high environmental persistence (Lim, 2021; Rahman et al., 2001; Wackett, 2021). These chemicals also raise a problem as they rapidly bioaccumulate in tissues of animals and human beings, owing to a high structural propensity to partition into lipid‐rich tissues (Desforges et al., 2018). Given the rising concern about their side effects, developing methods for the remediation of organohalides has become a critical research topic (He et al., 2021; Xu et al., 2023; Yao et al., 2021). Among the various remediation strategies, bioremediation of organohalides holds the potential to effectively remove organohalide pollutants in a cost‐effective and environmental‐friendly manner.

Bioremediation of organohalides requires functional microorganisms to achieve efficient degradation. However, the increasing diversity and amount of organohalides, as well as their emerging alternatives, represent new challenges. These challenges necessitate an expansion of the diversity of microorganisms capable of metabolizing these pollutants (Xu et al., 2023). For instance, there is limited information available on the biodegradation of per‐ and polyfluoroalkyl substances (PFASs), and ongoing efforts are required to identify microbes being capable of efficiently degrading and detoxifying these polyfluorinated pollutants (Jin et al., 2023; Wackett, 2022; Wackett, 2024; Yu et al., 2022). In contrast, recent advancements in methodology and technology have facilitated the isolation and cultivation of new microorganisms, providing new avenues for the discovery of functional microorganisms for bioremediation of organohalide pollution (Lewis et al., 2021; Liang et al., 2022). Additionally, the distinct characteristics of organohalides and their environmental context determine their transformation potential and consequently govern the involving functional microorganisms and underlying mechanisms. For example, recent research has indicated a shift in the major organohalide‐respiring bacteria (OHRB) responsible for organohalide dehalogenation upon niche transition from terrestrial to marine settings (Xu et al., 2023). Despite organohalide remediation being a topic of study for several decades, the growing variety of organohalides and new findings regarding their attenuation and associated microorganisms prompt rethinking bioconversion of organohalide pollutants, which may provide a roadmap for their future bioremediation applications.

WHY DO WE NEED TO PAY ATTENTION TO BIOREMEDIATION OF ORGANOHALIDE POLLUTION?

Compared to non‐halogenated organic compounds, the electronegative halogen atom in organohalides forms a polar carbon‐halogen bond, resulting in an electron‐deficient carbon atom. This bond also elevates the carbon's oxidation state, rendering organohalides resistant to oxidation processes. Consequently, organohalides accumulate ubiquitously in various oxidative environmental matrices at relatively high concentrations (Dorner et al., 2022; He et al., 2021; Lu et al., 2021). Their intrinsic stability necessitates reduction‐based remediation and associated sustainable management. A diverse range of organohalides are included in various priority pollutant control lists, such as the Stockholm Convention (Stockholm Convention on Persistent Organic Pollutants (SCPOP), 2024), the National Priority List by the USEPA (U.S. EPA, 2014), the Substance Priority List by the Agency for Toxic Substances and Disease Registry (ATSDR, 2022) and the Emerging Pollutant List by China's Ministry of Ecology and Environment (Ministry of Ecology and Environment (MEE), 2023). For example, out of 126 priority pollutants, 69 are organohalides, with 35 classified as persistent organic pollutants (POPs). The ATSDR ranks organohalides among the top 50 chemicals of concern for human health, with partial of them remaining on this list for decades (ATSDR, 2022). In China's emerging pollutant list, 13 out of the 14 groups comprise organohalides (Table 1), including short‐chain chloroalkanes and halogenated antibiotics (e.g. chlortetracycline, ciprofloxacin, enrofloxacin). Notably, the priority list of organohalides is keeping expanding with the emergence of their alternatives. Epidemiologically and experimentally, these halogenated chemicals are linked to various human diseases, including cancer (Abolhassani et al., 2019; Steenland & Winquist, 2021), immune deficiencies (Fletcher et al., 2019; Peinado et al., 2020) and neurodegeneration (Foguth et al., 2020; Saravi & Dehpour, 2016), with mitochondrial toxicity playing an important role. The direct interaction of organohalides with mitochondria underscores their potential to induce mitochondrial dysfunction and target molecular mechanisms within the mitochondria (Denslow & Martyniuk, 2023). Given these concerns, the effective remediation of these organohalide pollution is urgent.

TABLE 1.

Fourteen types of emerging pollutants are listed by the Ministry of Ecology and Environment of China.

No.	Classification	Compounds	CAS	Formula	Molecular weight (g/mol)	Chemical structure	LD50 (mg/kg)	Usage
1	Perfluorooctane sulfonic acid, its salts and Perfluorooctane‐sulfonyl fluoride	Perfluorooctane sulfonic acid (PFOS)	1763‐23‐1	C8HF17O3S	500.13		579	Surfactant
		Perfluorooctanesulfonyl fluoride	307‐35‐7	C8F18O2S	502.12		/	Surfactant/anti‐inflammatory
		Potassium heptadecafluoro‐1‐octanesulfonate	2795‐39‐3	C8F17KO3S	538.22		/	Surfactant/anti‐inflammatory/chromium mist inhibitor
2	Perfluorooctanoic acid, its salts and related compounds	Perfluorooctanoic acid (PFOA)	335‐67‐1	C8H4F15O2	414.07		Intraperitoneal LD50: 189	Surfactant/dispersant
		Sodium perfluorooctanoate	335‐95‐5	C8F15NaO2	436.05		/	Surfactant
		Potassium perfluorooctanoate	2395‐00‐8	C8F15KO2	452.16		/	Surfactant
3	Decabrominated diphenyl ether	Decabrominated diphenyl ether (Deca‐BDE)	1163‐19‐5	C12Br10O	959.17		500	Flame retardant
4	Short‐Chain Chlorinated Paraffins, SCCPs	Chloroalkanes C10‐13	85535‐84‐8	/	/		Intraperitoneal LD50: 4000	Flame retardant
		Chloroalkanes C6‐18	68920‐70‐7	/	/		/	Flame retardant
		Chloroalkanes C12‐13	71011‐12‐6	/	/		/	Flame retardant
5	Hexachloro‐1,3‐butadiene	Hexachloro‐1,3‐butadiene (HCBD)	87‐68‐3	C4Cl6	260.76		82	Solvent/heat exchanger agent
6	Pentachlorophenol, its salts and esters	Pentachlorophenol (PCP)	87‐86‐5	C6HCl5O	266.34		27	Antimicrobials
		Sodium pentachlorophenolate (Na‐PCP)	131‐52‐2	C6Cl5NaO	288.32		126	Herbicides
		Pentachlorophenyl dodecanoate	3772‐94‐9	C18H23Cl5O2	448.64		/	Preservative
7	Dicofol	Dicofol	115‐32‐2	C14H9Cl5O	370.49		575	Insecticide
		o, p′‐Dicofol	10606‐46‐9	C14H9Cl5O	370.49		/	Insecticide
8	Perfluorohexane sulfonic acid, its salts and its related compounds	Perfluorohexanesulfonic acid (PFHxS)	355‐46‐4	C6HF13O3S	400.12		/	Surfactant/metal plating
		Sodium perfluorohexanesulfonate	82382‐12‐5	C6F13NaO3S	424.11		/	Waterproofing agent/oil‐proofing agent
		Potassium perflurohexanesulfonate	3871‐99‐6	C6F13KO3S	438.20		/	Surfactant/metal plating
9	Dechlorane Plus and its syn‐isomer and anti‐isomer	Dechlorane Plus (DPs)	13560‐89‐9	C18H12Cl12	653.72		25,000	Flame retardant
		Dechlorane Plus syn	135821‐03‐3	C18H12Cl12	653.72		/	Flame retardant
		Dechlorane Plus anti	135821‐74‐8	C18H12Cl12	653.72		/	Flame retardant
10	Dichloromethane	Dichloromethane (DCM)	75‐09‐2	CH2Cl2	84.93		1600	Solvent/foaming agent
11	Trichloromethane	Trichloromethane (TCM)	67‐66‐3	CHCl3	119.38		908	Solvent/anaesthetic
12	Antibiotics	Chlortetracycline (CTC)	57‐62‐5	C22H23ClN2O8	478.88		1500	Antibiotics
		Ciprofloxacin (CIP)	85721‐33‐1	C17H18FN3O3	331.34		>2000	Antibiotics
		Enrofloxacin (ENR)	93106‐60‐6	C19H22FN3O3	359.40		5000	Antibiotics
13	Banned POPs	Hexabromocyclododecane (HBCD)	25637‐99‐4	C12H18Br6	641.70		/	Flame retardant
		Chlordane	57‐74‐9	C10H6Cl8	409.78		200	Insecticide
		Mirex	2385‐85‐5	C10Cl12	545.54		235	Insecticide
		Hexachlorobenzene (HCB)	118‐74‐1	C6Cl6	284.78		3500	Disinfectant
		Dichlorodiphenyltrichloroethane (DDT)	50‐29‐3	C14H9Cl5	354.49		113	Insecticide
		α‐Hexachlorocyclohexane (α‐HCH)	319‐84‐6	C6H6Cl6	290.83		177	Insecticide
		β‐Hexachlorocyclohexane (β‐HCH)	319‐85‐7	C6H6Cl6	290.83		6000	Insecticide
		Lindane	58‐89‐9	C6H6Cl6	290.83		88~270	Insecticide
		Endosulfan	115‐29‐7	C9H6Cl6O3S	406.93		40~50	Insecticide
		Polychlorinated biphenyls (PCBs)	/	C12H(10−n)Cln	/		PCB3: 4250; PCB4: 11000; PCB5: 1295	Flame retardant
14	Nonylphenol	Nonylphenol (4‐NP)	25154‐52‐3	C15H24O	220.35		1620	Surfactant/plasticizer/lubricants/disinfectant
		Nonylphenol (NP)	84852‐15‐3	C15H24O	220.35		1300	Surfactant/plasticizer/lubricants/disinfectant

Note: LD₅₀ data were retraced from Chemical Toxicity Database, 2024 (https://www.drugfuture.com/toxic/search.aspx) and ChinaChemNet, 2024 (https://cheman.chemnet.com/dict/msds.html).

Pioneering field monitoring in the 1970s revealed widespread contamination of underground water and soil with organohalides. This was followed by laboratory validation of bioremediation techniques during the 1980s to 2000s, initial field‐scale applications in the 2000s, and the establishment of mature bioremediation technologies thereafter. A breakthrough in bioremediation of organohalide pollution was the discovery of Dehalococcoides being capable of completely detoxifying chloroethenes, which dechlorinate tetrachloroethylene (PCE) to non‐toxic ethene (He et al., 2003; Maymo‐Gatell et al., 1997). To date, many OHRB‐based in situ bioremdiation cases have been successfully implicated for removing chloroethene pollutants (Aulenta et al., 2007; Pérez‐de‐Mora et al., 2014; Scheutz et al., 2008; Wu, Wang, et al., 2023; Xiao et al., 2020), although these processes often require years or even longer remediation time. By contrast, despite extensive laboratory research on aromatic organohalides, such as chlorophenols, chlorobenzenes, PCBs and PBDEs (Payne et al., 2011, 2017; Su et al., 2019; Wang et al., 2014), reports on their field‐scale bioremediation applications remain scarce (Pakdeesusuk et al., 2005). The main obstacle to the implementation of organohalide bioremediation is the lack of effective strategies for their extensive attenuation and even complete mineralization, which awaits further investigation.

HOW DO MICROORGANISMS CONVERT PERSISTENT ORGANOHALIDES?

The transformation of organohalides in environmental matrices can be categorized into two major lines: abiotic and biotic transformation. Abiotic transformation, occurring in matrices such as soil and sediment, is primarily mediated by iron‐/sulfur‐containing minerals with reducing equivalents. Nonetheless, the efficiency of natural abiotic transformation is significantly lower than that of biotic transformation (Darlington et al., 2008; Liu et al., 2023; McCarty, 1996). As a result, microbial attenuation plays a major role in natural conversion of organohalide pollutants. For example, around 59% of ocean‐derived CH₃Cl is transformed through microbial processes (Keppler et al., 2020). The biotic conversion of organohalides in natural environments could be grouped into reductive (dehalogenation) and oxidative processes (oxidative degradation).

Reductive processes

Highly halogenated organohalides, owing to their strong electrophilicity and high redox potential, need to undergo a reductive dehalogenation process to be transformed into less halogenated organics prior to their further oxidative degradation (Figure 1). Concurrently, these resistant organohalides tend to accumulate in anoxic/anaerobic environments (soil and sediment), where OHRB primarily facilitate their reductive dehalogenation (Adrian & Löffler, 2016). In the context of reductive dehalogenation mediated by organohalide‐respiring bacteria, electrons are transferred from H₂ or organic compounds to organohalides via membrane‐associated electron transport chains (Löffler & Edwards, 2006; Mohn & Tiedje, 1992; Wang et al., 2018). This reductive process may be coupled with oxidative phosphorylation, serving as an energy‐yielding respiratory process for cell growth of OHRB (Dehalococcoides, Dehalogenimonas and Dehalobacter as obligate OHRB and Sulfurospirillum, Geobacter and Desulfitobacterium as non‐obligate OHRB) (Fincker & Spormann, 2017). Notably, protons and electrons are derived directly from H₂ or NADH through dehydrogenases. Menaquinones or a complex iron–sulfur molybdoenzyme subunit then act as central electron carriers for non‐obligate and obligate OHRB, respectively (Cimmino et al., 2023; Kublik et al., 2016; Schubert et al., 2018). Ultimately, reductive dehalogenases (RDases) function as terminal reductases in the respiratory electron transport chain and play a central role in catalysing halogen removal from organohalides (Jugder et al., 2015; Wang et al., 2018). To date, crystal structures of reductive dehalogenases, such as PceA from Sulfurospirillum multivorans (Bommer et al., 2014), NpRdhA from Nitratreductor pacificus pht‐3B (Payne et al., 2015) and PceA from Desulfitobacterium hafniense TCE1 (Cimmino et al., 2023), have been elucidated. These enzymes share several structural commonalities, with a core domain that harbours the corrinoid cofactor, cobalamin, across various dehalogenases (Cimmino et al., 2023; Fincker & Spormann, 2017). Changes in the cobamides structure can affect RDases function (Schubert et al., 2019). Based on existing biochemical knowledge, particularly cobamide biochemistry, it is generally assumed that the reactive species at an RDase's active site is cob(I)alamin (Parthasarathy et al., 2015; Schubert et al., 2019; Yan et al., 2018). Three distinct mechanisms have been proposed to explain cobalamin‐dependent halogen elimination (Fincker & Spormann, 2017), all of which suggest organohalides with high redox potentials (E₀′ = 300–500 mV) to be readily reduced by Co(I) (E₀′_Co(II)/Co(I) = −350 mV) (Schumacher et al., 1997; van de Pas et al., 1999). These observations will explain why OHRB are obligate anaerobes and why reductive dehalogenation predominantly occurs under anaerobic/anoxic conditions. One exception is the glutathione‐dependent dehalogenation in aerobic bacteria Flavobacterium (Xun et al., 1992), of which underlying mechanisms remain explored. Hence, highly halogenated organohalides undergo reductive dehalogenation prior to their subsequent attenuation.

FIGURE 1.

Biotransformation pathways of chloroethenes through reductive dehalogenation and oxidative degradation.

Oxidative processes

A wide array of phylogenetically diverse aerobic microorganisms, including Bacillus, Comamonas, Pseudomonas, Rhodococcus, Paraburkholderia and Sphingomonas, have been identified for their capability in conversion of lowly halogenated organohalides via oxidative degradation (Bako et al., 2021; Han et al., 2023; Hara et al., 2022; Paliya et al., 2021; Wittich et al., 1992). The enzymes facilitating the aerobic metabolism or co‐metabolism of organohalides include monooxygenase, dioxygenase, hydrolase, haloacid dehalogenase and oxidative dehalogenase (Ito et al., 2017; Keim et al., 1999; Liu et al., 1995; Pimviriyakul et al., 2018). Enzymatic reactions are arranged through debranching ring breakage, dehalogenation/hydroxylation and ring‐cleavage oxidation/reduction to achieve mineralization (Pimviriyakul et al., 2020). A notable example of an aerobic bacterium being capable of degrading highly toxic polychlorinated dibenzo‐p‐dioxins/furans is Sphingomonas wittichii RW1. This microorganism employs a dioxygenase to mediate angular ring hydroxylation of various chlorinated dioxins and furans, resulting in the formation of trihydroxylated diphenylethers/trihydroxybiphenyls, predominantly through the dioxygenolytic cleavage of the ortho‐hydroxylated aromatic rings (Keim et al., 1999; Wittich et al., 1992). Although this strain effectively mineralizes non‐chlorinated dioxins and furans, the breakdown of chlorinated congeners is impeded by steric hindrances, with the rates of dioxygenase‐mediated transformations decreasing along with the increasing degree of chlorination. This trend of reduced aerobic bacterial transformation rates with increased number of halogen substitution is also observed for other organohalides such as PCBs and PBDEs (Furukawa et al., 1979; Han et al., 2023; Robrock et al., 2009). For instance, the catabolic activity of Alcaligenes sp. Y42 towards various trichlorobiphenyl congeners produces specific metabolites, yet highly chlorinated biphenyls remain unaltered (Furukawa et al., 1979). The effectiveness of aerobic degradation is limited to lowly halogenated organohalides, specifically those with one to four halogens (Han et al., 2023). The challenge in oxidative degradation of organohalides is further exacerbated by their increased hydrophobicity and specific thermodynamic properties. The halogenation extent positively correlates with the hydrophobicity of organohalides (Jiao & Li, 2010), and hydrophobic compounds typically induce stress in microbial cells, reducing biomolecular interactions (Bhaganna et al., 2010). Furthermore, the alignment of theoretically thermodynamic properties of organohalide congeners with observed oxidation rates suggests that halogenation‐induced changes in inherent thermodynamics significantly affect dioxygenase‐catalysing oxidation activities (Nam et al., 2014). The encouraging aspect is that microbial oxidative degradation leads to the mineralization of organohalides, showcasing its potential as a promising biotechnological approach for the attenuation or even complete mineralization of these pollutants.

Hybrid processes for extensive bioremediation of organohalide pollution

Microbial transformation of organohalides plays a crucial role in both natural attenuation and engineered bioremediation, as microorganisms utilize the energy derived from redox reactions to support their cell growth (Kracke et al., 2015). The Gibbs free energy (ΔG) available from each electron transport chain depends on the redox potential difference (ΔE) across all coupled reactions between the electron donors and acceptors. For example, the electrochemical potential difference (ΔE⁰′) for the reductive dehalogenation of organohalides, with hydrogen (H₂, E⁰′_H+/H2 = −414 mV) serving as the typical upstream electron donor and organohalides (E⁰′ typically ranging from 300 to 500 mV) as the downstream electron acceptor, exceeds 700 mV (Figure 2). This process theoretically generates 2–3 ATP/Cl⁻ via a chemiosmotic mechanism (Dolfing, 2000; Dolfing & Novak, 2015; Holliger et al., 1998). Consequently, delineating the interplay between the redox potential of organohalide and microbial processes emerges as a critical endeavour for augmenting the activity of microorganisms and, by extension, the bioremediation efficiency of organohalides. The redox potential of organic pollutants indicates their ability to act as either electron donors or acceptors in microbial respiration. Organic pollutants with a high redox potential (e.g. highly halogenated organohalides with E⁰′ from 300 to 500 mV) are suitable electron acceptors for reduction processes, such as microbial reductive dehalogenation (Wang et al., 2018), whereas those with a low redox potential (e.g. lowly halogenated organohalides or non‐halogenated organics, such as E⁰′_Benzoate/CO2 = −280 mV) can donate electrons in oxidation processes, with O₂, nitrate, sulfate or other compounds as the electron acceptors (Figure 2) (Bouwer & Zehnder, 1993). Meanwhile, the environmental redox potential, referred to as the Eh value, measures the environment's tendency to either donate or accept electrons, indicating whether oxidative (positive Eh values) or reductive conditions (negative Eh values) prevail. This distinction aligns with microbial oxidative or reductive processes, respectively (Hook et al., 1988). Therefore, the success of remediation efforts depends on aligning the pollutant's redox potential with the environmental redox capacity, underscoring the importance of carefully managing environmental conditions during remediation. As demonstrated in previous laboratory studies (Long et al., 2015; Payne et al., 2017; Wang et al., 2022), the complete mineralization of highly halogenated organohalides requires coupling OHRB‐based reductive dehalogenation and subsequent oxidative degradation (Figure 1). For instance, a two‐stage treatment combining microbial reductive dechlorination and oxidative degradation has been applied to remediate highly chlorinated PCBs, with PCB180 showing a removal efficiency of 37.1% through reductive dechlorination and subsequent aerobic microcosms degrading 48.2% of the less chlorinated PCB through oxidative processes (Wang et al., 2022). Nonetheless, fluctuations in redox potential may pose stress to microbes and directly affect the reactivity and bioavailability of organohalides, thus influencing microbial degradation efficiency (Wu et al., 2019). For example, studies have observed relatively low removal efficiencies of organohalides, potentially due to the inhibition of aerobic bacterial activity under prolonged anaerobic conditions (Pijuan et al., 2009). Therefore, while theoretically energy‐efficient and thermodynamically favourable, the sequential reductive‐oxidative process for bioremediation of organohalide pollution underscores the ongoing need to discover or design environmental settings for enhancing the practical application of this biotechnological method.

FIGURE 2.

Redox ladder and half‐reactions showing the transformation of typically oxidized organohalides and reduced organics as oxidation–reduction potential decline. Redox processes at standard conditions (i.e. pH 7 and 25°C).

In addition to reductive‐oxidative processes, several other hybrid processes have shown to be promising for complete detoxification of organohalides. For example, Jácome and colleagues demonstrated the biodegradation of lindane (the γ isomer of hexachlorocyclohexane) into non‐toxic end products (i.e. CO₂ and CH₄) under reduced conditions (Puentes Jácome et al., 2021). They utilized three anaerobic cultures to accomplish the dechlorination of lindane to monochlorobenzene, monochlorobenzene to benzene, and methanogenic benzene degradation, respectively. In the first two cultures, distinct populations of OHRB (Dehalobacter) were responsible for the dechlorination processes, while benzene‐degrading anaerobes (such as Deltaproteobacterial candidate Sva0485), along with fermenters, acetogens and methanogens in the third culture, were involved in converting benzene to CO₂ and CH₄ (Luo et al., 2016). This experimental evidence suggests that reductive‐fermenting processes could also achieve the complete detoxification of organohalides, particularly in anoxic environments where adjusting site redox potential is challenging. Another recent interesting finding to achieve such processes is possibly due to the potential multifunctionality of Dehalobacter (Chen, Fisch, et al., 2020). As an obligate OHRB, Dehalobacter was proposed to transform dichloromethane through both dechlorination and fermentation (Bulka et al., 2023; Chen, Tong, et al., 2020). Bulka and colleagues present a Dehalobacter‐containing enrichment culture that can simultaneously metabolize chloroform and dichloromethane (Bulka et al., 2023). The culture was maintained for over 1400 days without the addition of an exogenous electron donor, indicating that dichloromethane metabolism could produce sufficient reducing equivalents (likely hydrogen) for chloroform reduction. This ‘self‐feeding’ could prolong bioremediation activity long after donor addition ends, and also reduce the cost of bioremediation.

In conclusion, depending on the redox potential of organohalides‐polluted sites and halogenation extent of these organohalides, four different strategies could be employed to enhance bioremediation efficiency (Figure 3):

1. Solely reductive processes can be utilized to completely detoxify highly halogenated organohalides in anaerobic conditions;

2. Reductive‐fermentative hybrid processes are suitable for degrading highly halogenated organohalides in anaerobic conditions;

3. Reductive‐oxidative hybrid processes are effective for degrading highly halogenated organohalides in redox‐changing conditions;

4. Oxidative processes can be employed to degrade lowly halogenated organohalides in aerobic conditions.

FIGURE 3.

Proposed redox‐potential‐mediated hybrid process to completely detoxify typical organohalides.

BIOREMEDIATION OF ORGANOHALIDE POLLUTION IN VARIED SCENARIOS

Contaminated sites

The accidental and deliberate release of large quantities of organohalide results in site contamination (Stringer & Johnston, 2001). Contaminated sites, such as aquatic sediments, submerged soils and groundwater, are usually oxygen‐depleted. Therefore, anaerobic bacteria capable of organohalide respiration have been of great importance as candidates for bioremediation (Lee & Manefield, 2023; Maphosa et al., 2012). Thus far, bioremediation of organohalides by employing OHRB, whether by stimulating indigenous organisms or by adding exogenous cultures, is the most common and successful remedial approach for restoring contaminated sites. For instance, several OHRB‐containing consortia, such as KB‐1™ (Duhamel & Edwards, 2006), Pinellas culture (Ellis et al., 2000) and Bachman‐aquifer enrichment (Lendvay et al., 2003), have already been commercialized and used for cleaning up organohalide‐contaminated sites. Most of the field studies mainly focus on the bioremediation of chloroethenes, such as PCE, trichloroethylene (TCE) and vinyl chloride (VC), which can be completely dechlorinated to non‐harmful ethene (Pérez‐de‐Mora et al., 2014; Scheutz et al., 2008; Wu, Shen, et al., 2023). However, for other organohalides, especially highly halogenated organics (e.g. PCBs), sole dehalogenating bioremediation often results in incomplete dehalogenation (Payne et al., 2011), necessitating other approaches (e.g. zero‐valent iron reduction, persulfate oxidation and aerobic biodegradation) to achieve complete detoxification (Wang et al., 2016, 2022; Wu & Wang, 2022).

Landfills

Landfills serve as an important reservoir of organohalides, primarily due to the disposal of organohalides‐containing solid waste. Studies have documented elevated concentrations of organohalide pollutants in landfill leachates, including chloroethenes, chlorobenzene, chlorophenoxypropionic acid, chlorophenols, PCBs and PBDEs (Grøn et al., 2000; Weber et al., 2011), posing potential risks to downstream aquifers and biota (Haarstad & Borch, 2008). Landfills typically exhibit anoxic conditions due to the presence of massive organic matter, along with rich sources of electron donors and electron shuttles (e.g. humic acids), which might facilitate organohalide respiration (Qin et al., 2022). Moreover, the methanogenic conditions in landfills enrich fermenters, acetogens and methanogens, which may be suitable for applying the reductive‐fermentative process for organohalide detoxification. However, a significant concern is the high levels of ammonium in landfills, which may inhibit the activity of OHRB and other functional microbes (Delgado et al., 2016). Studies have shown that the rate of dehalogenation decreases by approximately 90% in the situation of ammonium concentrations exceeding 2 g/L (Delgado et al., 2016). Therefore, achieving high removal efficiency of organohalide pollutants in landfills requires selecting OHRB and associated degraders being tolerant to high concentrations of ammonium for the bioaugmentation process.

Paddy soil

Paddy soil represents a typical human‐affected wetland globally and is vulnerable to pollution from various sources, including organohalides from rapidly growing industrial zones, fertilizer application, organic solid waste and wastewater irrigation (Akram et al., 2018; Luo et al., 2023). Paddy cultivation involves distinct agricultural practices, including cyclic wetting and drying, seasonal ploughing and leaving fields fallow, leading to frequent redox cycles between anoxic and oxic conditions (Kögel‐Knabner et al., 2010). These redox oscillations induced by paddy management profoundly impact the structure and functionality of microbial communities, thereby affecting short‐term biogeochemical cycles. Considering the organohalide contamination and natural redox variations in paddy fields, these ecosystems are optimally predisposed to employing the reductive‐oxidative hybrid process for bioremediation of organohalide pollution. Laboratory experiments have investigated hypoxic bioremediation of synthetic paddy soils contaminated by various organohalides, such as octachlorinated dibenzodioxin and dibenzofuran (Wu et al., 2019), TBBPA (Liu et al., 2013), and PCBs (Chen et al., 2014). Nevertheless, comprehension of the redox influences on the hypoxic decomposition of organohalides in paddy soils and the associated microbial dynamics remains limited. Compared to aquatic environments, paddy soils feature abundant environmental buffers, like redox‐altering natural organic matter and iron/manganese oxides (Li et al., 2015). These elements are crucial in modulating the activities and efficiencies of microbial populations involved in bioremediation processes (Akram et al., 2018). By capitalizing on this aspect, incorporating redox mediators into paddy soils, including organic matter, biochar and nanoscale materials, has the potential to significantly enhance the bioremediation efficiency.

Wastewater treatment plants (WWTPs)

Wastewater treatment plants (WWTPs) serve as a primary sink for organohalides that can enter wastewater streams through various ways, including the disposal of cleaning fluids by manufacturing facilities, disposal of consumer products, landfill leachate and precipitation (Gottschall et al., 2017; Regnery & Püttmann, 2010; Xu et al., 2021). Traditional WWTPs are not specifically designed to remove organohalides as micropollutants. Consequently, only a minor to moderate fraction of organohalides entering WWTPs are converted during treatment, with the majority being adsorbed onto sludge and the remainder being discharged with the effluent (Margot et al., 2015; Ricklund et al., 2009). Conventional WWTPs typically incorporate sequenced anaerobic and aerobic processes, primarily aiming to remove organic matter and nutrients (Mishra et al., 2023), thereby providing potentially ideal niches for the application of the reductive‐oxidative process for removing organohalides. Nonetheless, the relatively low removal efficiency for organohalides in WWTPs may be attributed to the short hydraulic retention time and sludge retention time, which is insufficient for the proliferation of organohalides‐converting bacteria, particularly the OHRB with extended generation times (Adrian & Löffler, 2016). However, increasing evidence have shown a high abundance of obligate OHRB in both anoxic and aerobic zones of WWTPs, indicating their capability for the reductive degradation of organohalides (Xu et al., 2024; Zhao et al., 2020). Beyond the limited duration of wastewater treatment systems, the process of anaerobic sludge digestion necessitates a much longer retention time to facilitate reductive dehalogenation (Smith et al., 2015). In addition, major proportion of organohalides in waste stream will absorb to sludge and enter sludge digesters. Bioaugmentation with OHRB in anaerobic sludge digesters will be a promising strategy to detoxify the organohalides in WWTPs. Then, a subsequent microbial process, such as aerobic composting, is essential for further elimination of dehalogenated products (Ponza et al., 2010). Therefore, reductive‐oxidative bioremediation could be integrated into upgraded WWTPs through the bioaugmentation of functional OHRB and aerobic degraders to remove organohalides from wastewater streams.

Treatment wetlands

Treatment wetlands (TWs), also known as constructed wetlands, are engineered systems designed specifically for the removal of pollutants from contaminated water (Faulwetter et al., 2009). Recent advancements in anoxic/oxic TWs have proven their effectiveness in the removal of nutrients and organic matter from wastewater (Wang et al., 2023). Unlike traditional WWTPs, TWs generally require a longer hydraulic retention time, typically between 2 and 10 days (Wu, Wang, et al., 2023), thus providing a suitable niche for the organohalides‐converting microorganisms. A diverse range of OHRB, including Dehalococcoides, Desulfitobacterium and Sulfurospirillum, have been identified in the constructed wetlands, indicating these environments may be favourable for their proliferation (Atashgahi et al., 2016). Research on the bioremediation of pharmaceuticals and personal care products, such as diclofenac and sulfonamides within anoxic–oxic TWs, has demonstrated substantial removal efficiency (Chen, Fisch, et al., 2020; Sochacki et al., 2018). These investigations have uncovered a variety of microbial metabolisms, leading to hydroxylated, methylated, dechlorinated and oxidative by‐products in the effluents from TWs (Sochacki et al., 2018). Nonetheless, specific underlying degradation mechanisms, functional microorganisms/genes and the interactions among different microbial populations remain to be fully elucidated. In addition to microbial mediation, plants within TWs play a crucial role in the transformation of pollutants (Bhatia & Goyal, 2014). Therefore, the complexity of remediation systems in TWs requires further investigation, especially the synergistic interactions between microbes and plants, highlighting the need for a focused approach to hypoxic biotechnology applications within these systems.

CHALLENGES, OPPORTUNITIES AND FUTURE PERSPECTIVES

Bioremediation stands as a promising biotechnology for removing organohalide contamination across diverse environmental matrices. The selection of a suitable bioremediation strategy hinges upon the redox properties of contaminated sites and halogenation extent of organohalides. Reductive processes, exemplified by reductive dehalogenation, are effective for removing highly halogenated organohalides with elevated redox potential. However, the sluggish growth of OHRB commonly leads to incomplete dehalogenation, resulting in the accumulation of lowly halogenated compounds. Consequently, the remediation of lowly halogenated organohalides with low redox potential typically necessitates an oxidative process. Thus, the adoption of appropriate bioremediation strategies needs to align with the specific characteristics of the organohalides to ensure sustainable management. Nevertheless, the implementation of such strategies should carefully consider the site‐specific conditions, encompassing the redox potential of environmental matrices, the presence of functional and indigenous microorganisms, and other technical factors influencing the bioremediation processes. To further augment the remediation efficiency, the following topics await further investigation:

Exploration and implication of microbial resources

The continuously expanding list of OHRB for microbial reductive dehalogenation of organohalides prompts the question of whether most organohalides can be effectively dehalogenated and detoxified by known and unknown OHRB. For example, the microorganisms for efficient defluorination of polyfluorinated pollutants remain largely unexplored, despite thermodynamic calculations suggesting that these processes are biologically feasible (Parsons et al., 2008). To date, only limited aerobic and anaerobic strains or microcosms have been identified to be capable of degrading PFASs, and their transformation mechanisms remain unclear (Jin et al., 2023; Yu et al., 2022). However, as the fastest‐evolving cellular form of life, prokaryotes may naturally evolve their catabolic capability to degrade polyfluorinated compounds (Wackett, 2021, 2024). One potential solution is to enrich and isolate new OHRB from extreme environments (Wackett, 2023). For instance, recent studies have discovered different OHRB species with novel RDases from marine or high‐salinity environments, such as Dehalobium (Xu et al., 2023) and Desulfoluna (Peng et al., 2020). Another intriguing aspect of this research is the indication that the multi‐functionality of RDases in OHRB is merely coincidental. In addition, many OHRBs have been found to utilize a single RDase for dehalogenating multiple organohalides (Qiu et al., 2020; Xu et al., 2022; Zhao et al., 2020), although the detailed mechanisms remain elusive. Therefore, further research is warranted to explore novel microbial resources (both microorganisms and functional enzymes) for remediation purposes, which will also advance our understanding of the fate and transport of organohalides in natural environments.

Coupling biotic and abiotic processes

Bioconversion holds promise for remediating organohalide‐contaminated environments, yet its application poses technical challenges due to several following factors: (i) the potential accumulation of toxic intermediates at high concentrations resulting from incomplete dehalogenation or transformation and (ii) the absence or low abundance of functional microbes catalysing the desired bioconversion process at bioremediation sites. To achieve effective remediation of organohalide pollution, complementary abiotic treatment processes may be employed to further detoxify them. For example, the integration of nanoscale zero‐valent iron (nZVI) with microbial reductive dehalogenation (Bio‐RD) has been proposed and implemented to achieve rapid and complete dehalogenation of organohalides (Kocur et al., 2016; Koenig et al., 2016; Wang et al., 2016). Another approach is to combine Bio‐RD with persulfate activation and oxidation (PAO) for highly halogenated organohalides that cannot be effectively remediated by Bio‐RD or PAO alone (Wu, Shen, et al., 2023). It is important to note that most abiotic treatments involve different materials/chemical agents, such as ZVI, metal–organic frameworks (MOFs) and biochar (Giannakoudakis et al., 2022), and their cytotoxicity to microorganisms, mobility and redox potential warrant further investigation.

Synthetic biology

Bioremediation of organohalide pollution typically involves multiple syntrophic groups of microorganisms, rather than the sole key functional population, such as syntrophic interactions among OHRB, fermenters and methanogens (Dehalococcoides‐Desulfovibrio‐Methanosarcina) in reductive dehalogenating microcosms (Men et al., 2012; Wang et al., 2019). In addition to the syntrophic microbes, the presence of different degrading microorganisms associated with co‐existing organohalides must be considered when developing synthetic microbial consortia (Jayaramaiah et al., 2022). Therefore, rather than solely relying on functional microbes, scientists are increasingly turning to synthetic biology to combine and adapt the most useful biological traits into bespoke microcosms for bioremediation (Rylott & Bruce, 2020). Several key research directions should focus on elucidating the interactions between syntrophic and functional microbes, as well as among different functional microbes, and understanding the dynamics of the degrading microbial community and indigenous microorganisms.

Ecological principles‐guided strategies

The successful bioremediation of organohalide pollution through bioaugmentation and biostimulation approaches mainly depends on the colonization of functional microorganisms. Developing strategies for establishing niches appropriate for the functional microorganisms requires consideration of the environmental habitats, microbial behaviour and interactions between indigenous and inoculated microorganisms. Recently, leveraging two ecological principles, that is, Priority Effects and Co‐existence Theory, three strategies, namely niche preparation, staggered fermentation and increased inoculation, were successfully implemented to enhance the niche colonization of OHRB in field bioremediation of chloroethene pollution (Wu, Wang, et al., 2023). This approach led to a significant increase in the abundance of inoculated OHRB, culminating in the complete dechlorination of TCE and VC to ethene within 5 and 3 months, respectively. Thus, employing interventions guided by ecological principles represents a promising avenue to expand the potential for in situ bioremediation of organohalide pollution.

Bioremediation of unconventional organohalide pollution

Unconventional organohalides constitute a subset of emerging contaminants, representing chemicals that have garnered increased attention due to their potential environmental and human health impacts. These compounds were initially developed as alternatives to banned organohalides (He et al., 2021). For example, PBDEs, once widely used as brominated flame retardants due to their prevalence, underwent a phase‐out from production and use between 2004 and 2013 (Sharkey et al., 2020). However, this phase‐out led to the emergence of alternative flame retardants, such as short‐chain chloroalkanes, Dechlorane Plus and TBBPA. Yet, it has become increasingly evident that some of these alternatives are also ubiquitous in the environmental matrices and pose ecological and health risks, prompting their inclusion in the Stockholm Convention and relevant emerging pollutant lists. Therefore, particular attention should be devoted to these alternatives, not only in environmental assessments but also in efforts towards their bioremediation.

P450 enzymes

Cytochrome P450 enzymes (P450s) are highly versatile redox enzymes that have attracted considerable interest for their potential applications in bioremediation. To date, over 300,000 cytochrome P450 sequences have been identified across diverse biota, including microorganisms, plants and mammals (Nelson, 2018). In addition to their well‐known role in the oxidative conversion of organohalides (Isin & Guengerich, 2007), P450s have also been found to catalyse reductive dehalogenation in the absence of oxygen (Behrendorff, 2021). For example, several P450 enzymes (e.g. 2B6, 2C9, 2E1, 3A4 from Rattus norvegicus; 6G1 from Drosophila melanogaster) have been shown to catalyse the reductive dechlorination of DDT, producing dichlorodiphenyldichloroethane (Joußen et al., 2008; Kitamura et al., 2002). The conformation of active sites, the orientation and redox properties of substrates relative to the heme group are crucial factors determining whether reductive dehalogenation by P450s will occur. Given their natural aerotolerance, P450s offer promising opportunities for improvement and deployment as reductive biocatalysts for refractory organohalides. Unlike OHRB‐derived reductive dehalogenases that are typically highly sensitive to oxygen, P450s can be deployed with higher flexibility. Furthermore, the diversity of P450s from various sources, including microbes, plants and metazoans, suggests potential for cooperative interactions, which warrant further investigation.

Bioinspired design

As mentioned, bioremediation is generally time‐consuming and may not completely degrade organohalides. To address the gap between biological approaches and the need for sustainable and effective environmental remediation, biochemists have begun synthesizing functional enzymes to directly target pollutants, thereby overcoming the inherent inefficiencies associated with microbial processes (Kuddus et al., 2024). In particular, atomically dispersed single‐metal‐site materials have emerged as promising platforms to meet these demands, owing to their uniformity of isolated active sites and tunable coordination environments through the host structure (Baek et al., 2018). Recently, Min and colleagues developed a single‐atomic‐site Co catalyst supported by carbon‐doped boron nitride (BCN) with locally polarized B–N bonds (Co SAs/BCN) to mimic the function of reductive dehalogenases (Min et al., 2021). The Co SAs/BCN catalyst exhibited high activity, achieving nearly complete dechlorination (~98%) of chloramphenicol, indicating its potential for the sustainable conversion of organohalides. These findings suggest that bioinspired design of novel materials may offer new opportunities for the development of sustainable and effective remediation strategies.

AUTHOR CONTRIBUTIONS

Qihong Lu: Conceptualization; project administration; writing – original draft; writing – review and editing; funding acquisition; visualization. Qi Liang: Methodology; data curation; visualization. Shanquan Wang: Conceptualization; writing – original draft; writing – review and editing; project administration; supervision; funding acquisition.

CONFLICT OF INTEREST STATEMENT

The authors declare no competing interest.

Lu, Q. , Liang, Q. & Wang, S. (2024) Burning question: Rethinking organohalide degradation strategy for bioremediation applications. Microbial Biotechnology, 17, e14539. Available from: 10.1111/1751-7915.14539