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Review

The Degradation of Polycyclic Aromatic Hydrocarbons by Biological Electrochemical System: A Mini-Review

by
Yu Tian
1,
Rumeng Wang
2,
Min Ji
2,
Ruimin Tian
1,
Renjie Wang
1,
Bo Zhang
3,
Shaopo Wang
1,* and
Lingjie Liu
1,2,*
1
School of Environmental and Municipal Engineering, Tianjin Chengjian University, Jinjing Road 26, Tianjin 300384, China
2
School of Environmental Science and Engineering, Tianjin University, Tianjin 300350, China
3
Tianjin Eco-City Water Investment and Construction Co., Ltd., Hexu Road 276, Tianjin 300467, China
*
Authors to whom correspondence should be addressed.
Water 2024, 16(17), 2424; https://doi.org/10.3390/w16172424
Submission received: 12 July 2024 / Revised: 28 July 2024 / Accepted: 30 July 2024 / Published: 28 August 2024
(This article belongs to the Special Issue Application of Electrochemical Technology for Water and Wastewater Treatment)
Browse Figures
Versions Notes

Abstract

:
Polycyclic aromatic hydrocarbons (PAHs) are persistent environmental pollutants commonly found in water and sediments, posing significant health risks due to their toxicity, carcinogenicity, and mutagenicity. The stable and sustainable degradation of PAHs has garnered significant attention from researchers. Biological electrochemical systems (BESs) offer a promising approach with advantages in energy efficiency, safety, environmental protection, and long-term operation. This review examines the degradation performance and microbial community dynamics of BESs in the treatment of PAH-contaminated water and sediments. Additionally, the metabolites formed during the degradation process were also summarized. This review summarizes the degradation characteristics of PAH-contaminated water and sediments and aims to guide future research and optimize BESs for effective remediation of PAHs in various environmental settings.

1. Introduction

Polycyclic aromatic hydrocarbons (PAHs) are a class of aromatic compounds with two or more benzene rings, characterized as persistent and refractory organic pollutants [1,2]. The primary sources of PAH emissions are anthropogenic activities, such as fossil fuel combustion, motor vehicles, and waste incineration [1,3,4,5]. PAHs are widespread and persistent in the environment, with high concentrations often detected in aquatic environments, including water and sediment [6,7]. Humans are frequently exposed to PAHs through water, with exposure pathways including ingestion, inhalation, and skin contact [8]. This exposure poses significant health risks, including carcinogenicity, mutagenicity, neurotoxicity, and immunotoxicity [2,9,10,11]. Notably, PAHs are difficult to degrade and persist in the environment, prompting numerous studies focused on their degradation.
Currently, the methods commonly used to remove PAHs from water include physical, chemical, and biological approaches. Physical methods primarily involve biochar adsorption of PAHs [12]. Chemical methods encompass advanced oxidation processes, photochemical degradation, and ultrasonic degradation. Among these, biological treatment is widely utilized, leveraging various bacterial, fungal, and algal species with the potential to degrade PAHs [13]. Biological treatment offers advantages such as environmental friendliness, low cost, and sustainability [14]. The efficiency of microbial degradation of PAHs largely depends on various culture conditions, including temperature, pH, nutrient availability, microbial population, chemical properties of PAHs, oxygen levels, and acclimatization [6]. However, biological treatment faces challenges such as low microbial activity and slow growth rates.
In recent years, biological electrochemical systems (BESs) have garnered extensive attention as a promising method for treating PAHs. BESs are a green remediation technology regulated by electrochemically active microorganisms on electrodes within the system [15]. This technology boasts advantages such as energy savings and sustainability [16]. In addition, integrating novel functional materials, particularly metal-organic frameworks (MOFs), with BESs could enhance PAH degradation, making this a promising area for further research and development. For example, Gd-doped Zinc-MOFs, Zn(II)-MOFs, Cu(II)-based MOFs, and Cd(II)-MOFs proved effective in degrading pollutants like antibiotics and dyes through photocatalytic and luminescent properties [17,18,19,20]. The biodegradation of PAHs has been extensively reviewed, covering microbial degradation mechanisms, bioremediation technologies and their applications, and the role of molecular biology in elucidating these mechanisms [6,14,21,22]. However, systematic summaries of the role and mechanism of biological electrochemical systems (BESs) in PAH degradation remain relatively scarce.
The main objectives of this review are to: (1) introduce microbial electrochemical systems; (2) summarize the removal characteristics of PAHs in water and sediments, including microbial characteristics; and (3) identify the metabolites of PAHs after treatment by BESs. This review aims to provide a systematic summary of the role and mechanisms of BESs in the degradation of PAHs, offering specific insights for future research and practical applications.

2. Overview of Biological Electrochemical Systems

BESs utilize microorganisms as electrochemical catalysts to convert chemical energy into electrical energy or valuable chemicals through electron transfer [23].

2.1. System Types of BES

BESs encompass various types, including microbial fuel cells (MFCs), microbial electrolysis cells (MECs), microbial desalination cells (MDCs), microbial electrosynthesis (MES), and other configurations [24].
The mechanisms of various types of BESs differ significantly. MFCs generate electricity by microbial catabolism of substrates, enhancing electron transfer and organic matter degradation through anode-cathode circuits [25]. MECs produce hydrogen gas by applying an external voltage to facilitate reduction reactions at the cathode, while microorganisms at the anode oxidize organic substrates [26]. MDCs desalinate water and generate electricity simultaneously by incorporating ion exchange membranes, driven by microbial oxidation at the anode and the resulting ion movement [27]. This dual functionality of water desalination and energy production makes MDCs a promising technology for addressing water scarcity and energy demands simultaneously [28]. MES converts CO2 into value-added chemicals using electric current, with autotrophic microorganisms at the cathode fixing CO2 into organic compounds, supporting sustainable biochemical production and carbon capture [29,30].

2.2. Electron Transfer in BES

Electron transfer in BESs is the key factor in achieving pollutant degradation [31,32]. In a BES, electron transfer occurs primarily through direct and indirect pathways [33]. Direct electron transfer involves the direct movement of electrons from microorganisms to the electrode, facilitated by conductive structures like pili or cytochromes [34,35]. Indirect electron transfer involves the use of mediators, which are compounds that shuttle electrons between the microorganisms and the electrode [33]. Electron mediators, such as flavins or quinones, enable electron transfer over greater distances or in less conductive environments, enhancing system efficiency [36]. Thus, optimizing these electron transfer pathways is crucial for enhancing the performance of microbial electrochemical systems, as they directly impact the efficiency of energy conversion and PAH degradation [37].

3. Electrochemical Bioremediation Technology for PAHs-Contaminated Sediments

3.1. Degradation of PAHs by BES

Sediment biological electrochemical systems (SBESs) showed significant potential in degrading PAHs in sediments, as shown in Table 1. SBESs could effectively degrade PAHs under both anaerobic and aerobic conditions. Under anaerobic conditions, the degradation rates of naphthalene, acenaphthene, and phenanthrene were 76.9%, 52.5%, and 36.8%, respectively. Similarly, effective degradation of naphthalene (41.7%), acenaphthene (31.4%), and phenanthrene (36.2%) was achieved under aerobic conditions [38]. The variation in these removal rates may be attributed to the fact that the terminal electron acceptors under anaerobic conditions, such as nitrates and sulfates, can promote the degradation of PAHs [38]. Moreover, a recent study demonstrated that carbon nanomaterial-modified electrodes significantly enhanced PAH degradation in sediment MFCs, with removal rates for phenanthrene and pyrene reaching up to 78.1% and 69.6%, respectively [39]. This improvement was attributed to increased electrochemically active surface areas and microbial diversity on the modified electrodes [39]. Furthermore, the addition of amorphous ferric hydroxide to the sediment increased the removal rates of phenanthrene and pyrene to 99.5% and 94.8%, respectively, within 240 days [40].
The degradation of PAHs by SBESs is limited by mass transfer efficiency [41]. Plant-driven mass transfer could improve the removal rates of phenanthrene and pyrene by 63.0% and 57.0%, respectively, after 82 days, which was 1.5 to 2.0 times higher than the rates observed in unplanted systems [41]. The synergistic effect of Acorus tatarinowii Schott and sediment MFCs significantly increased the degradation rates of pyrene and benzo[a]pyrene to 87.18% and 76.40%, respectively, within 367 days. This represented an approximately 60% improvement over natural degradation [42].
Appropriate auxiliary substrate supplementation could maintain or improve the performance of MFCs and promote the degradation of PAHs [25]. The study of Wu et al. [43] showed that glucose accelerated the bioelectrochemical degradation of phenanthrene in sediment, increasing its removal rate by 1.09 times. Additionally, adding 5–10 mg/L starch could improve the removal rate of PAHs, as starch served as a co-metabolic substrate for sediment MFCs, enhancing PAH removal [44]. Unlike glucose, starch was partially hydrolyzed into low molecular weight organic acids under anaerobic conditions [45]. These organic acids could be used by anode microorganisms as carbon sources and energy sources, thereby increasing the bioavailability of PAHs and adjusting the pH to improve microbial communities [44].
Table 1. Performance of MFCs in the degradation of PAHs in sediments.
Table 1. Performance of MFCs in the degradation of PAHs in sediments.
Sediment SourceConditions; Temperature; PeriodPAH Type; ConcentrationDegradation Efficiency (%)NotesReferences
Macritchie reservoirAerobic; 27 °C; 45 dNaphthalene;
50 ppm		50 [38]
Aerobic; 27 °C; 45 dAcenaphthene;
50 ppm		31.4
Aerobic; 27 °C; 45 dPhenanthrene;
50 ppm		36.2
Anaerobic; 27 °C; 45 dNaphthalene;
50 ppm		76
Anaerobic; 27 °C; 45 dAcenaphthene;
50 ppm		52.5
Anaerobic; 27 °C; 45 dPhenanthrene;
50 ppm		36.8
A large shallow lakeAnaerobic; 25 °C; 240 dPhenanthrene;
10 mg/kg		96.14 [40]
Anaerobic; 25 °C; 240 dPyrene;
5 mg/kg		92.13
Anaerobic; 25 °C; 240 dPhenanthrene;
10 mg/kg		99.47Amorphous ferric hydroxide
(16 g wet weight)
Anaerobic; 25 °C; 240 dPyrene;
5 mg/kg		94.79
Urban river surface sedimentsAnaerobic; 24 °C; 82 dPhenanthrene;
1.38 mg/kg		62.98 [41]
Anaerobic; 24 °C; 82 dPyrene;
1.28 mg/kg		57.02
Lake surface sedimentsAnaerobic; 367 dPyrene;
1.28 mg/kg		55.73 [46]
Anaerobic; 367 dBenzo[a]pyrene;
1.28 mg/kg		47.2
Anaerobic; 367 dPyrene;
1.28 mg/kg		87.18Macrophyte Acorus Calamus
Anaerobic; 367 dBenzo[a]pyrene;
1.28 mg/kg		76.4
Aquaculture pond sedimentAnaerobic; 25 °C; 68 dNaphthalene;
39.4–43.1 mg/kg		39.2 [44]
Anaerobic; 25 °C; 68 dAcenaphthene;
50.3–52.3 mg/kg		23.4
Anaerobic; 25 °C; 68 dPyrene;
50.4–53.6 mg/kg		19.1
Anaerobic; 25 °C; 68 dNaphthalene69.9Starch (10 mg/g)
Anaerobic; 25 °C; 68 dAcenaphthene55.6
Anaerobic; 25 °C; 68 dPyrene46.8

3.2. Functional Microbial Communities

Firmicutes, Bacteroidetes, and Proteobacteria were common bacterial phyla in the anodic environment of sediment MFCs, as shown in Figure 1 [41]. Firmicutes have the ability to decompose complex organic matter, participate in the transfer of extracellular electrons, and exhibit electrochemical activity [47]. The higher relative abundance of Bacteroidetes was related to the higher PAH removal performance of microbial fuel cells [44]. Some researchers speculated that Bacteroidetes could contribute to PAH removal [44]. In addition, some non-dominant bacteria were also involved in substrate consumption and electron transfer [44]. Although these bacteria were not directly related to the degradation of PAHs, they created a synergistic environment to enhance the degradation of PAHs [48]. For example, Acidobacteria focused on the decomposition of polymers and the metabolism of carbon compounds [49]. Acidobacteria could provide additional carbon sources and nutrients, which supported the metabolic activities of PAH-degrading bacteria, further facilitating PAH degradation [50,51]. Co-occurrence network analysis showed that Acidobacteria played an important role in the microbial interaction network in PAH-contaminated environments [52]. In the process of starch as a co-metabolism substrate involved in the removal of PAHs, some fermentation bacteria in Anaerolineaceae could convert some organic matter into short-chain fatty acids [53]. These short-chain fatty acids were then utilized by anodic microorganisms to promote the removal of PAHs [41].
At the genus level, multiple microbial genera such as Pseudomonas, Diaphorobacter, Thauera, Lysobacter, and others showed significant contributions in the PAH degradation process through multiple pathways [39,43]. Pseudomonas could promote the degradation of phenanthrene during the iron-based biochar-enhanced BES process [43]. In addition, Zhang et al. [54] found that adding nitrate to the sediment could promote the growth of Pseudomonas, thereby enhancing the biodegradation effect of SMFC on various PAHs (phenanthrene, pyrene, and benzo[a]pyrene). Pseudomonas could shuttle compounds by releasing electrons, such as phenazine and pyocyanin [55]. These compounds promote electron transfer to the anode, enhancing the microbial co-oxidation of PAHs [56]. Moreover, Pseudomonas could directly participate in the degradation of phenanthrene by secreting PAH degradation-related enzymes, such as PAH dioxygenase and catechol-2,3-dioxygenase, which perform ring dihydroxylation and ring-opening reactions [57]. Pseudomonas might also increase the solubility and improve the bioavailability of phenanthrene by producing biosurfactants, thereby promoting the biodegradation of phenanthrene [58,59].
Moreover, the abundance of Diaphorobacter, Thauera, and Lysobacter genera was positively correlated with the degradation of phenanthrene, indicating a potential role in the biodegradation process of PAHs [39]. Among them, Diaphorobacter could efficiently degrade phenanthrene through the phthalic acid pathway [60]. Thauera and Lysobacter had the function of degrading various PAHs such as naphthalene, phenanthrene, and pyrene [61,62,63]. Additionally, Diphenylbutadiene might initiate the oxidation of anthracene and fluorene and collaborate with other strains to metabolize various PAHs [64].
The development of BESs that leverage functional microbial communities presents a promising solution for the efficient degradation of polycyclic aromatic hydrocarbons (PAHs) in contaminated sediments. Future research should concentrate on elucidating the synergistic interactions between dominant and non-dominant bacterial phyla to enhance PAH degradation efficiency further. Additionally, exploring the metabolic pathways and electron transfer mechanisms of key genera such as Pseudomonas, Diaphorobacter, Thauera, and Lysobacter would provide deeper insights into optimizing BESs for environmental remediation.

4. Electrochemical Bioremediation Technology for PAHs Contaminated Water

PAHs in water primarily originate from industrial discharges, urban runoff, and incomplete combustion of organic matter [65]. MFCs and MECs have emerged as effective green biotechnologies for treating PAH-contaminated water [66]. Currently, MFCs and MECs are used to degrade various PAHs, including phenanthrene, anthracene, naphthalene, pyrene, and benzo[a]pyrene (Figure 2a).

4.1. Degradation of PAHs by MFC

MFC showed significant potential in the degradation of PAHs in water bodies, as shown in Table 2. The bioavailability of phenanthrene, a kind of polycyclic aromatic hydrocarbon commonly found in water environments, was higher than other PAHs due to its relatively low molecular weight [67]. Adelaja et al. [68] used MFC to degrade phenanthrene, and the degradation rate was more than 97%. In addition, Hua et al. [69] used a single-chamber air cathode MFC and sodium acetate co-metabolism to effectively improve the effective removal of high concentration (20 mg/L) phenanthrene (98.44%). This effect was due to the fact that microorganisms destroyed the structure of phenanthrene through anaerobic metabolism and established tolerance to phenanthrene. Recently, researchers found that loading catalysts on the electrode surface could effectively improve the degradation of phenanthrene. For example, Huang et al. [70] used a 1.5CuCo@NC-800 catalyst to treat the cathode, and achieved a phenanthrene removal rate of more than 98% at a concentration of 1–10 mg/L. The addition of the catalyst improved the efficiency of mass transfer and electron transfer, and thus improved the performance of the electrocatalytic oxygen reduction reaction, as shown in Figure 2b. The high graphitization and large specific surface area of the catalyst were the key factors in promoting the removal of phenanthrene [70].
In addition, researchers also found other common PAHs such as anthracene, naphthalene, pyrene, and benzo[a]pyrene could be effectively degraded by MFCs. Wang et al. [71] used nano-zero-valent iron (nZVI) as an anode material to successfully remove anthracene within 182 days, and the removal rate reached 96.4%. The use of nZVI could increase the easily available metabolic substrates around the anode, thereby improving the treatment efficiency [72,73]. Gambino et al. [74] treated 80 mg/L of naphthalene and achieved a removal rate of 97.13% within 35 days. Pyrene was a common PAH. For 30 mg/L pyrene, the degradation rate could reach 88.1% [75]. In addition, some studies confirmed that benzo[a]pyrene in the range of 0.17–20 mg/L could also be effectively degraded by MFCs [74,76].
Table 2. Performance of MFCs in the degradation of PAHs in simulated wastewater.
Table 2. Performance of MFCs in the degradation of PAHs in simulated wastewater.
PAHs Type; ConcentrationConditions; Temperature; PeriodDegradation Efficiency (%)NotesReferences
Phenanthrene;
5 mg/L		Anaerobic; 30 °C; 40 h98.84 [70]
Phenanthrene;
10 mg/L		Anaerobic; 30 °C; 40 h98.77
Phenanthrene;
5 mg/L		Anaerobic; 30 °C; 40 h95.3CuCo@NC 800
(3 mg/cm2)
Phenanthrene;
10 mg/L		Anaerobic; 30 °C; 40 h98.37
Phenanthrene;
5 mg/L		Anaerobic; 30 °C; 60 h97.05 [69]
Phenanthrene;
10 mg/L		Anaerobic; 30 °C; 70 h94.9
Phenanthrene;
20 mg/L		Anaerobic; 30 °C; 80 h98.44
Phenanthrene
0.17 mg/L		Anaerobic; 26 °C; 48 h93.6 [71]
Anthracene;
0.17 mg/L		Anaerobic; 26 °C; 48 h95.3
Phenanthrene;
0.17 mg/L		Anaerobic; 26 °C; 48 h95.8nZVI modified carbon fiber felt electrode
Anthracene;
0.17 mg/L		Anaerobic; 26 °C; 48 h96.5
Pyrene;
5 mg/L		Anaerobic; 30 °C; 48 h44.8 [75]
Pyrene;
20 mg/L		Anaerobic; 30 °C; 48 h72.4
Pyrene;
30 mg/L		Anaerobic; 30 °C; 48 h88.1

4.2. Degradation of PAHs by MECs

Unlike MFCs, MECs apply an external voltage to enhance extracellular electron transfer and thereby increase microbial electrochemical activity [77].
The cathode reaction kinetics was more favorable, thereby promoting the degradation of refractory organic matter at the cathode by increasing the potential difference [77]. Min et al. [66] found that the removal rate of naphthalene by MECs reached 94.5% using nickel foam as the cathode material. Additionally, Ding et al. [78] demonstrated that MECs could achieve an 85.15% degradation efficiency for naphthalene even under NaCl concentrations as high as 60 g/L. This was achieved by maintaining microbial activity through increased biofilm thickness and a higher live/dead bacteria ratio.
Another important approach to utilizing MECs for the degradation of PAHs is the addition of co-metabolized substrates. Ding et al. [79] demonstrated that using residual phenol in coal chemical wastewater as a co-metabolism substrate could promote the degradation efficiency of naphthalene more effectively than commonly used sodium acetate and glucose, achieving a removal rate of 84.11% within 5 days. This improvement was attributed to the presence of phenol, which significantly enhanced the expression of the ncrA gene mediating the reduction of the benzene ring in naphthalene and the nmsA gene mediating the decarboxylation of naphthalene, thereby facilitating more effective degradation [79]. Interestingly, although the addition of sodium acetate and glucose diversified the naphthalene degradation pathways, phenol as a co-metabolite substrate led to a more complete degradation of naphthalene [79].

4.3. Functional Microbial Communities

The synergistic existence of microorganisms in BESs could effectively degrade polycyclic aromatic hydrocarbons (PAHs) in water, e.g., Proteobacteria, Firmicutes, and Actinobacteria at the phylum level, as well as Pseudomonas, Bacillus, Streptomyces, and Geobacter at the genus level.
Several key microbial phyla played significant roles in this degradation process. Proteobacteria and Firmicutes exhibited excellent PAH degradation capabilities in both aquatic and sediment environments [41,71]. Proteobacteria degraded complex organic compounds through various metabolic pathways, such as acylglycerol degradation, the Arnon–Buchanan cycle, and the Calvin cycle [80,81]. Firmicutes participated in the oxidative degradation of PAHs by secreting exogenous enzymes such as catalases and dioxygenases [82]. Actinobacteria, known for their strong ability to degrade complex organic compounds, actively degraded PAHs [7]. The study of Behera et al. [83] demonstrated that several genera belonging to the phylum Actinobacteria, such as Streptomyces, Rhodococcus, Microbacterium, and Arthrobacter, could degrade PAHs through modified in situ cleavage pathways.
At the genus level, several microbial genera presented significant capabilities in degrading PAHs. Pseudomonas was frequently found in PAH-contaminated waters [84]. Pseudomonas facilitated PAH degradation by producing biosurfactants, electron transfer proteins, and exogenous enzymes, as shown in Figure 2c [68]. Bacillus and Streptomyces primarily degraded PAHs by secreting various degradative enzymes. Bacillus secreted catalases [71,85], while Streptomyces effectively degraded PAHs in aquatic environments by secreting lignin peroxidases and monooxygenase [86]. Additionally, biosurfactants played a crucial role in the biodegradation of PAHs, aiding in the dissolution of PAHs and thus increasing their bioavailability [87]. For example, Streptomyces rochei, known for producing biosurfactants, effectively degraded PAHs containing 3–4 rings [88]. Geobacter was a special functional bacterium that was widely present in microbial electrochemical systems [89]. Geobacter mediated extracellular electron transfer and enhanced PAH degradation efficiency through electricity generation processes, as shown in Figure 2c [79]. Geobacter transferred electrons to electrodes, driving the degradation of complex organic compounds while generating electrical energy [90]. Their conductive pili and cytochromes acted as biological wires, facilitating long-distance electron transfer [91]. The presence of Geobacter in BESs significantly boosted PAH breakdown by creating a conducive environment for electron transfer and energy production [90].
With the continuous development of microbial detection technology, more types of PAH-degrading bacteria were discovered. In the future, the dynamic changes of microbial communities and their synergistic effects under different environmental conditions should be further explored to optimize the degradation effect in practical applications.

5. Metabolite Analysis

In understanding the effectiveness of BESs in degrading PAHs in different environments, it is crucial to analyze the metabolites formed during the degradation of PAHs.
The biodegradation process of phenanthrene mainly involved three main pathways, i.e., hydroxylation, methylation, and carboxylation [92]. Firstly, phenanthrene underwent the hydroxylation reaction to form intermediate products such as p-cresol. These intermediates were then gradually converted into smaller molecules such as phenol through methylation and carboxylation reactions [93]. Finally, these smaller molecules were completely degraded and converted into CO2 through microbial metabolism.
The anaerobic biodegradation of naphthalene first occurred through the carboxylation process. Naphthalene was converted to 2-naphthoic acid by carboxylase [92]. Subsequently, 2-naphthoic acid was further converted to 2-naphthoyl-CoA, which involved a series of enzymatic reactions [94]. The ring structure of 2-naphthoic acid remained unchanged during the metabolic process until it was reduced to 5,6-dihydronaphthalene-2-carboxylic acid and finally converted to CO2 [95].
In the biodegradation of pyrene, the initial step involved the production of 1-hydroxypyrene by monooxygenase and dioxygenase enzymes [75]. Then, this intermediate was transformed into smaller molecules, such as 1-hydroxy-2-naphthalic acid, in different ways. 1-Hydroxy-2-naphthoic acid was further degraded to produce small molecular compounds such as phenylacetic acid and terephthalic acid. These compounds were eventually utilized by microorganisms and participated in the TCA cycle [75,96].
Furthermore, co-metabolism would play an increasingly important role in the biodegradation of PAHs. This mechanism involved using non-target pollutants, such as growth substrates, to induce the production of key enzymes or cofactors required to degrade target pollutants, offering a promising approach for enhancing PAH degradation efficiency [97]. Specifically, using phenol as a co-metabolism substrate has been shown to improve the degradation efficiency of naphthalene, promote biomass accumulation, and enhance the metabolic activity of microorganisms [79]. Moreover, the addition of co-metabolism substrates significantly increased the expression of ncrA and nmsA genes, which were closely related to the key enzyme activities in the anaerobic degradation of naphthalene (Figure 2b) [79]. Overall, co-metabolism offered a valuable strategy for enhancing PAH biodegradation in BESs.

6. Conclusions and Prospects

BESs presented a promising approach for the degradation of PAHs in both water and sediment environments. Compared to conventional biodegradation methods, BESs offer significant advantages, including energy efficiency, safety, environmental friendliness, and sustained operation. The effectiveness of BESs in degrading PAHs was influenced by various factors such as substrate concentration, applied voltage, electrode distance, and the composition of the microbial community involved. Moreover, lower molecular weight PAHs were degraded more readily and extensively compared to higher molecular weight PAHs. Enhancements to the traditional BES setup, including the addition of catalysts, surfactants, and iron modifications, had demonstrated improved degradation efficiencies. Furthermore, co-metabolism showed significant potential in enhancing PAH degradation by using non-target pollutants to induce the production of key enzymes or cofactors required for the degradation of target pollutants. Additionally, photocatalysis has shown potential in enhancing the degradation of various pollutants, such as trichlorophenol and antibiotics, in MFCs. This suggests a promising direction for future enhancements in the degradation of PAHs.
Future research should focus on the integration of molecular biology techniques to further elucidate the mechanisms underlying PAH degradation in BESs. Prospects for advancing this field include: (1) exploring the combination of photocatalysts with BESs to promote the production of active substances, thereby potentially improving the degradation efficiency of PAHs; (2) investigating the role of gene expression and enzyme activity in the biodegradation process to provide deeper insights into optimizing BESs for environmental remediation; (3) exploring the synergistic interactions within microbial communities and their metabolic pathways to enhance the overall efficiency of PAH degradation; and (4) developing and testing strategies for the intensification of PAH degradation by BESs, such as exogenous addition of substances and co-metabolism of substrates, to improve degradation efficiency and sustainability. These efforts will contribute to the effective remediation of PAH-contaminated environments and the advancement of sustainable bioremediation technologies.

Author Contributions

Y.T.: Investigation, Data curation, Formal analysis, Writing the original draft. R.W. (Rumeng Wang): Conceptualization, Writing—review and editing. M.J.: Validation, Visualization. R.T.: Resources, Writing—review and editing. R.W. (Renjie Wang): Resources, Writing—review and editing. B.Z.: Resources, Writing—review and editing. S.W.: Writing—review and editing. L.L.: Methodology, Supervision; Writing—review and editing. All authors have read and agreed to the published version of the manuscript.

Funding

The research leading to these results has received funding from the National Natural Science Foundation of China (No. 52070141), the National Natural Science Foundation of China (No. 42307178), the Key Project of the Tianjin Natural Science Foundation (No. 22JCQNJC00020), and the Graduate Scientific Research Innovation Project of Tianjin (No. 2022SKYZ175).

Conflicts of Interest

Author Bo Zhang was employed by the company Tianjin Eco-City Water Investment and Construction Co., Ltd. The remaining authors declare that the research was conducted in the absence of any commercial or financial relationships that could be construed as a potential conflict of interest.

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Water 16 02424 g001
Figure 1. Electrochemical bioremediation technology for PAH-contaminated sediments. (a) Degradation of PAHs by SBES under both anaerobic and aerobic conditions. (b) Enhancement of PAH degradation by plant-driven mass transfer. (c) Enhancement of PAH degradation by starch and glucose addition in SBESs. (d) Functional microbial communities in SBESs.
Figure 1. Electrochemical bioremediation technology for PAH-contaminated sediments. (a) Degradation of PAHs by SBES under both anaerobic and aerobic conditions. (b) Enhancement of PAH degradation by plant-driven mass transfer. (c) Enhancement of PAH degradation by starch and glucose addition in SBESs. (d) Functional microbial communities in SBESs.
Water 16 02424 g001
Water 16 02424 g002
Figure 2. Electrochemical bioremediation technology for PAHs contaminated water. (a) Schematic diagram of the BES for PAH degradation in water. (b) Strategies to enhance the performance of BESs in degrading PAHs. (c) Mechanisms of PAH degradation by different microbial genera. The red arrows in the figure indicate promotion.
Figure 2. Electrochemical bioremediation technology for PAHs contaminated water. (a) Schematic diagram of the BES for PAH degradation in water. (b) Strategies to enhance the performance of BESs in degrading PAHs. (c) Mechanisms of PAH degradation by different microbial genera. The red arrows in the figure indicate promotion.
Water 16 02424 g002
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Tian, Y.; Wang, R.; Ji, M.; Tian, R.; Wang, R.; Zhang, B.; Wang, S.; Liu, L. The Degradation of Polycyclic Aromatic Hydrocarbons by Biological Electrochemical System: A Mini-Review. Water 2024, 16, 2424. https://doi.org/10.3390/w16172424

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Tian Y, Wang R, Ji M, Tian R, Wang R, Zhang B, Wang S, Liu L. The Degradation of Polycyclic Aromatic Hydrocarbons by Biological Electrochemical System: A Mini-Review. Water. 2024; 16(17):2424. https://doi.org/10.3390/w16172424

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Tian, Yu, Rumeng Wang, Min Ji, Ruimin Tian, Renjie Wang, Bo Zhang, Shaopo Wang, and Lingjie Liu. 2024. "The Degradation of Polycyclic Aromatic Hydrocarbons by Biological Electrochemical System: A Mini-Review" Water 16, no. 17: 2424. https://doi.org/10.3390/w16172424

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