Three-Dimensional Biofilm Electrode Reactors with a Triple-Layer Particle Electrode for Highly Efficient Treatment of Micro-Polluted Water Sources

Abstract

Micro-polluted water, which is widespread in rural areas, poses a serious health risk. To address this issue, we propose a three-dimensional biofilm electrode reactor with triple-layer particle electrodes (TL-BERs) for the decentralized and small-scale treatment of micro-polluted water. The first and second layers of the electrode, granular activated carbon (GAC) and biological ceramsite (BC), respectively, are responsible for electric field oxidation and microbial degradation, respectively, while the third, quartz sand (QS), is responsible for further improving turbidity and pollutant removal. Our tests indicated that the TL-BER-treated effluent met the drinking water quality standards of China. At 10 V, the average turbidity, COD_Mn, NH₄⁺-N, and UV₂₅₄ removal rates of the TL-BERs system were 97.66%, 61.11%, 91.67%, and 72.94%, respectively. Furthermore, the intensities of the main fluorescence peaks, A, B, C, and D, of the raw water sample, decreased by 36.67%, 66.22%, 67.08%, and 69.76%, respectively, after treatment, indicating that tryptophan-like proteins, fulvic acid, and humic acid were also effectively removed. High-throughput sequencing analysis showed the enrichment of microorganisms, such as Proteobacteria, Bacteroidota, and Actinobacteriota, which play important roles in the removal of various pollutants. Therefore, the application of this strategy will enable the practical treatment of micro-polluted water.

1. Introduction

Rural drinking water sources, such as rivers, lakes, reservoirs, and rainwater, are generally little affected by industrial pollution; however, they are characterized by excessively high levels of ammonia nitrogen (NH₄⁺-N), turbidity, and organic pollutants [1,2,3]. For rural communities, especially remote areas in northwest China, constructing a centralized water supply system is challenging [4]. For example, on the Qinghai-Tibet Plateau and in other alpine pastoral areas, the local nomadic herdsmen drink water directly from nearby rivers, streams, and wells or perform simple filtration and sedimentation before drinking the water. Further, their living areas, which are often livestock gathering areas, can easily trigger the pollution of nearby water bodies, resulting in the intermittent availability of safe drinking water as well as periods during which certain pollutants exceed their standard levels. Except for a few clear water sources, most of the water sources in this area are slightly turbid and odorous. The consumption of such contaminated water can cause diseases, such as diarrhea, cholera, dysentery, typhoid, and polio [5]. Therefore, it is necessary to develop appropriate water treatment technologies that are characterized by low chemical addition as well as simple installation, operation, management, and maintenance, to ensure a steady supply of safe drinking water in such rural regions [6,7,8,9].

The conventional treatment process, flocculation-filtration-disinfection, cannot effectively remove the above-mentioned pollutants. Further, micro-polluted surface water is also often treated via adsorption, catalytic oxidation, and membrane filtration [10,11,12]. However, these approaches have a number of shortcomings: adsorption accumulates contaminants, but does not completely degrade micro-pollutants; catalytic oxidation is limited by the complexity of the associated technology; and membrane filtration is limited by the high cost of operation and fouling problems. Therefore, these approaches are not suitable for the decentralized treatment of drinking water in highland pastoral areas. Bio-slow sand filters (BSSFs) are economically feasible, easy to operate, and show high efficiency [13,14]. They are also suitable for decentralized water treatment [15]. Further, considered as one of the earliest known water treatment processes, BSSFs are suitable for the removal of NH₄⁺-N and organic pollutants, and for lowering turbidity [16,17]. For example, Zhao [18] used BSSFs to treat collected rainwater with a high chemical oxygen demand (COD), high NH₄⁺ concentration, and poor sensory indices (turbidity, color, visible matter, and smell), obtaining good results with respect to microbial pathogens. Liu [15] used BSSFs to treat rural drinking water and found that during the stable operation period, the effluent NH₄⁺-N content met the standards for drinking water quality in China (0.5 mg/L), whereas turbidity did not (1 NTU). Therefore, lowering turbidity and organic pollutant contents during water purification remains challenging. Notably, water sources in high-altitude pastoral areas are prone to low temperature, low turbidity, seasonal low temperature, and high turbidity water quality characteristics, which lead to low microbial activity on the surface of the filter media, resulting in poor water quality.

In recent years, the application of electrochemical technology in water treatment, in particular three-dimensional (3D) electrochemical technology, has received considerable attention owing to its ease of operation, small footprint, and short hydraulic retention time [19]. Specifically, 3D biofilm electrode reactors (3D-BERs), which integrate biofilm technology with 3D-ERs, are beneficial to treatment processes. Several mechanisms, including adsorption, biodegradation, electro-adsorption, electrochemical oxidation, and electro-biodegradation, contribute to pollutant degradation in 3D-BERs [20]. The accumulation of microorganisms induces biofouling on the electrode surface owing to the presence of particles on the electrodes. However, the synergy between electricity and microorganisms favors pollutant removal. For electro-biodegradation in particular, the synergy between electricity and microorganisms is critical for the removal of refractory pollutants. The associated mechanisms mainly include two aspects. First, enhanced biodegradation owing to the stimulation of microbial metabolism by the electric field and the acclimation of the microbial community [21], and second, the positive contribution of the intermediates generated from the electro-biodegradation process. Because of their lower operation costs compared with those of 3D-ERs, 3D-BERs have mainly been considered for the pretreatment of pharmaceutical wastewater [22]. For example, 3D-BERs were constructed to treat Rhodamine B (RhB) [20], where the application of voltage successfully promoted the degradation of RhB. Three processes, electro-adsorption, electrochemical oxidation, and electro-biodegradation, were identified as contributing to RhB degradation. Biodegradation of ibuprofen has also recently been demonstrated in a 3D-BER. The optimal conditions were identified at a current density of 12.73 A/m² and an HRT (hydraulic retention time) of 3.5 h [23]. In addition, Zhou [24] developed 3D-BERs for groundwater remediation using electrochemical and biotechnological techniques. To simultaneously improve denitrification performance and bring about the effective degradation of organic pollutants from micro-polluted groundwater, a novel 3D bio-electrochemical reactor with granular activated carbon (GAC), introduced into a traditional two-dimensional (2D) reactor as the third electrode, was developed. In this 3D reactor, the denitrification rate improved significantly to 0.288 mg NO₃–N/cm²/d, and the current efficiency reached 285%. Furthermore, the bioreactor demonstrated stable performance under variable conditions, indicating its potential for application in groundwater remediation. This 3D-BER technology has been primarily used for refractory wastewater treatment or denitrification. However, studies on its application to treat micro-polluted water have not yet been reported.

Therefore, in this study, we combined 3D-BERs with BSSFs to obtain a 3D-BER with triple-layer particle electrodes (TL-BERs) comprising three layers, namely, GAC, biological ceramsite (BC), and quartz sand (QS), to optimize the distribution of biological populations in the biofilm on the surface of the media based on the establishment of an electric field in the BSSFs; this improved the activity of the molecules or ions participating in the electrochemical reaction. Thus, the electric field oxidation and microbial degradation in the GAC and BC layers of the particle electrodes were maximized, and finally, the removal of turbidity and pollutants was further enhanced by the QS layer, resulting in efficient purification of micro-polluted water. The removal efficiencies of turbidity, COD_Mn, NH₄⁺-N, and UV₂₅₄ in the TL-BERs system were investigated and 3D fluorescence spectroscopy (EEMs) was performed to analyze the degradation of dissolved organic matter (DOM). Additionally, the microbial communities in the novel TL-BERs system and their functions were investigated. Therefore, the findings of this study may highlight a solution for the decentralized and small-scale treatment of micro-polluted water using novel TL-BERs.

2. Materials and Methods

2.1. Wastewater Source

As shown in Figure 1, the study area, located in Maqu County, Gansu Province, on the eastern part of the Qinghai-Tibet Plateau, is characterized by an average altitude of ~3500–3800 m, average annual temperature of 2.8 °C, and annual rainfall of ~633–782 mm. In this area, nomadic herdsmen directly drink water from nearby rivers, streams, and wells, or simply treat the water via filtration and sedimentation. Further, their living areas are often livestock gathering areas, which can easily trigger the pollution of nearby water sources, resulting in high levels of turbidity, NH₄⁺-N, COD_Mn, DOM, and other pollutants.

In this study, we simulated the water quality characteristics of alpine pastoral water and used this simulated water sample as a reference for our water distribution experiments. To simulate the micro-polluted water, alpine pasture grass (250 g), cow dung (35 g), and soil (50 g) from a nomadic pastoral area in Awancang town, Maqu County, were mixed with 10 L of tap water for 2 days to obtain a high concentration mother liquor, which was then mixed with tap water in the ratio 1:12 to obtain the experimental water samples. The water quality characteristics are shown in

Table 1. Characteristics of micro-polluted water.

Parameter	Turbidity /NTU	Temperature/°C	pH	Conductivity /µS·cm−1	DO /mg·L−1	NH4+-N /mg·L−1	UV254	CODMn /mg·L−1
Simulated water quality	~10–50	~1–20	~8.2–8.6	~250–350	~5–8	~0.07–0.87	~0.050–0.110	~2.5–10
Water quality	~9–25	~8–15	~8.4–8.6	~350–420	~5.8–7.8	~0.28–0.65	~0.055–0.095	~3.2–5.5 *

Note: * The experimental setup was equipped with a pretreatment device. The quality of the simulated water was less than that of the actual water from the alpine pastoral area.

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2.2. Experimental Setup and Start-Up

2.2.1. Experimental Setup

As shown in Figure 2, in this study, we used polymethyl methacrylate organic glass (length = 1900 mm; inner thickness [φ, inner diameter] = 150 mm) to design 3D biofilm electrode reactors with triple-layer particle electrodes (TL-BERs). Specifically, the overall design of the reactor comprised four parts: the water cover area, filter layer, support layer, and water collection layer. The filter layer comprised GAC (φ = 3 mm; thickness = 400 mm), BC (φ = 2–3 mm; thickness = 300 mm), and QS (φ = 0.5–1.0 mm; thickness = 400 mm). Regarding the composite gradation from top to bottom, the support layer comprised coarse sand (φ = 0.9–1.5 mm; thickness = 30 mm), gravel (φ = 3–5 mm; thickness = 50 mm), and pebbles (φ = 3–10 mm; thickness = 120 mm). Further, a titanium-based ruthenium dioxide mesh was placed on the outside surface of the filter layer (inner wall of the filter column) as the anode, while a pure titanium column was set up in the middle of the layer as the cathode, and wires connecting both electrodes were used to regulate the power supply between the electrodes.

The filtration speed of the TL-BERs was 0.2 m/h [25]. After the raw water was lifted to the water distributor by an electromagnetic diaphragm metering pump (EDM pump), it was discharged through a three-way water-level regulator after passing through the four parts of the system from top to bottom. The reactor was equipped with backwash water and air-flushing to clean the filter media. Further, the experiments were performed using an electromagnetic diaphragm metering pump (WS-09-03-S, Wilwofford, Beijing, China), DC-regulated power supply (LW303D-30V10A, Longway Power Supply, Hong Kong), backwash pump (12VDC/3A), and backwash air pump (AP-50A).

2.2.2. Experimental Procedure

The reactor adopted the natural membrane hanging method with a filtration rate of 0.2 m/h, and started the membrane hanging at a voltage of 0 V. Microorganisms in the influent water were trapped on the surface of the filter media, where they used the nutrients in the water to grow and reproduce continuously, eventually forming a stable biofilm; at the membrane hanging stage, water samples were taken from the inlet and outlet of the device every 3 days, each time at 10:00 and 16:00, and turbidity, COD_Mn, UV₂₅₄, NH₄⁺-N, dissolved oxygen (DO), pH, and conductivity were measured; the average value was considered as the measured index value for that day, until the pollution index value suddenly changed, after which samples were collected for analysis every other day until the membrane hanging experiment was considered successful, i.e., when the removal rate of COD_Mn was stable at >30% [26]. The reactor voltage was gradually increased from 0 to 25 V. After 10 d of successful film hook-up and stable operation, a 2.5 V charge was applied. With this increase in voltage, the pollutant removal effect became unstable owing to microbial adaptation to the new voltage. Thus, continuous sampling and measurement of water quality parameters were performed on the fifth day. After another 5 d, the voltage was increased to 5 V. This pattern of operation was continued, and the removal efficiencies of COD_Mn, NH₄⁺-N, turbidity, UV₂₅₄, pH, and DO were measured at 2.5, 5.0, 7.5, 10.0, 12.0, 15.0, 20.0, and 25.0 V. The optimal voltage was selected and the treatment efficacy at this voltage was compared with that at which the conventional BSSF operates (0 V) to analyze pollutant removal effects under two different working conditions, charged and uncharged. To fully explore the degradation effect of pollutants and analyze the pollutant removal mechanism in the different filter layers, the pollutant removal effect under the two different working conditions (charged and uncharged) was recorded and water samples were collected from different filter layers regularly. Finally, the filter media were subjected to high-throughput sequencing, and the microbial diversity on the surfaces of the filter media was analyzed in conjunction with the stratification mechanism to explore the microbial diversity in the GAC and BC layers and to validate the mechanism of the migration and conversion of organic matter from a microbial perspective.

2.3. Analytical Methods

COD_Mn and NH₄⁺-N were assessed using the potassium permanganate and Nascent reagent photometric methods, respectively, while UV₂₅₄ was assessed using the UV–visible photometric method (UVmini1240, Shimadzu, Japan). Turbidity was measured using a WGZ-1A turbidity meter (Xinrui, Shanghai, China), pH using a PHS-3C pH meter (Rex, Shanghai, China), DO using a HQ30d single channel digital meter (HACH, Loveland, CO, USA), and conductivity using a DDS-11A electrical conductivity meter (Rex).

To obtain 3D fluorescence maps, a fluorescence spectrophotometer (F-7100; Hitachi, Tokyo, Japan) was used. In this regard, to obtain a wide range of data encompassing all the organic compounds present, the scan ranges for analysis were set at 200–450 nm and 250–550 nm for excitation and emission, respectively [27]. The scan speed was fixed at 12,000 nm/min at a response time of 0.002 s and a PMT voltage of 700 V. Further, the slit width was fixed at 5 nm for both excitation and emission. Before each measurement, ultrapure water was used as the experimental blank, and during analysis, the blank was subtracted to deduct Raman scattering. The water samples were filtered through a 0.45-μm pore-size filter membrane to determine the 3D fluorescence spectrum, which was divided into five regions as previously described by Chen [27], who combined the fluorescence intensity change of the characteristic peak with the blue-shift phenomenon to reflect the relative transformation process of DOM in the 3D fluorescence spectrum region.

For high flux measurement, the filter media in the reactor were sampled at the optimum voltage of 10 V, and the fluxes were labeled as follows: AHA (GAC anode), AHC (GAC cathode), ATA (BC anode), and ATC (BC cathode). The filter media in the reactor were also sampled at 0 V, and the fluxes were labeled as follows: BH (GAC layer) and BT (BC layer). Biofilm samples (filter media) were collected during the stable operation of the device. At 10 V, filter media surrounding the titanium-based ruthenium dioxide mesh anode and pure titanium column cathode were collected, while at 0 V, GAC filter media and BC filter media were collected. These were placed in 50-mL sterile centrifuge tubes, sealed, and placed in an insulated box containing dry ice; all samples were collected in triplicate and sent to the testing company for DNA extraction. Biological samples were sampled and analyzed separately. Low-quality reads were removed and raw sequence data were optimized. Further, the optimized high-quality representative sequences were divided into operational taxonomic units (OTUs) with 97% sequence identity. Redundancy analysis (RDA) was also performed using the Mantel test to determine the relationship between microbial populations with relative abundance >1% at the phylum and genus levels and also in terms of physicochemical factors.

3. Results and Discussion

3.1. Pollutant Removal Effects under Different Voltages

In this study, the natural enrichment membrane hanging method was adopted, and the filtration speed and surface temperature of the filter were maintained at 0.2 m/h and ~8–15 °C, respectively, while the reactor voltage for the hanging membrane experiment was 0 V. After the biofilm was successfully established, voltages of 2.5, 5.0, 7.5, 10, 12, 15, 20, and 25 V were applied sequentially, and pollutant removal was measured after 4 d of stable operation at each voltage. The variations in each index at the different voltages and their removal rates are shown in Figure 3.

From Figure 3a, it is evident that with sequential increases in voltage, the pH of the effluent water decreased from 8.45 to 6.20, gradually changing from weakly alkaline to weakly acidic. This was possibly a result of the polar plate producing hydroxyl radicals, hydrogen peroxide, and other aqueous oxidants, which indirectly oxidized the organic matter in water, producing substances such as carboxylic acids and CO₂, which rendered the pH weakly acidic [28]. The DO of the effluent water increased from 4.45 to 8.56, higher than the DO of the influent water, which may be related to the fact that oxygen precipitation by the electrode plate increased with increasing voltage, resulting in the generation of a large amount of oxygen [29]. Further, the oxygen generated could not be used by microorganisms on time; thus, the DO content of the effluent water increased gradually. From Figure 3b, it is evident that at voltages below 10 V, the turbidity of the effluent was stable at 0.3 ± 0.2 NTU, as but as the voltage increased beyond 12 V, the turbidity increased markedly. This may be due to the fact that as the voltage increased, oxygen and hydrogen precipitation capacity increased, accompanied by the production of a large number of bubbles, resulting in the formation of a micro-suspension inside the filter media. Simultaneously, the resulting aeration strength established a large shear and frictional force on the surface of the filter media particles, resulting in the severe breakdown of the sludge flocs on the surface of the filter media, and biofilm loss [30]. Further, from Figure 3c, it is evident that NH₄⁺-N and COD_Mn both had higher removal rates at 2.5 and 10 V. This is because the low voltage primarily increased the activity of biological enzymes in microorganisms, promoting their metabolism and growth [31], and also increased the electron transfer rate [32]. The highest COD_Mn, NH₄⁺-N, and UV₂₅₄ removal rates of 61.11%, 91.67%, and 72.94%, respectively, were achieved at 10 V. Further, owing to the action of electricity, these were decomposed into readily degradable components, which were then further biodegraded on the biofilm attached to the filler, or by electrical stimulation, which enhanced the activity of the biofilms and, in turn, the biodegradability of pollutants [23]. When the voltage of the TL-BERs increased beyond 10 V, the removal rates of COD_Mn, NH₄⁺-N, and UV₂₅₄ decreased. Specifically, at 25 V, the average removal rates of these pollutants were 41.04%, 54.55%, and 38.75%, respectively. This can be explained by the dysregulation of microbial metabolism caused by the excessively high voltage, resulting in limited microbial degradation, hence microbial inactivation [33].

Analysis and discussion based on Figure 3 showed that a voltage of 10 V was optimal for the removal of COD_Mn, NH₄⁺-N, and UV₂₅₄. At this optimal voltage, the average removal rates of these pollutants were 61.11%, 91.67%, and 72.94%, respectively. These removal rates were 36.38%, 39.63%, and 38.20% higher, respectively, than the average year-on-year removal rates obtained at 0 V, suggesting that the removal of organic pollutants by the TL-BERs resulted from the synergistic action of multiple mechanisms, including biofilter adsorption, microbial degradation, and electric field oxidation [34]. Although power was consumed during this process, the power consumption at the optimal voltage of 10 V was 0.2 kW·h/m³, which is acceptable in terms of energy consumption for drinking water treatment. Therefore, to investigate the efficiency of the system, a voltage of 10 V was used in further investigations of the removal mechanism of pollutants between each filter layer.

3.2. Pollutant Removal Effects of Different Filter Layers

As shown in Figure 4b, the average removal rates of NH₄⁺-N, COD_Mn, and UV₂₅₄ in the GAC layer at working voltages of 0 and 10 V were 18.43%, 24.68%, and 30.77%, and 59.67%, 14.81%, and −75.00%, respectively. This indicated that the removal of NH₄⁺-N was mainly concentrated in the GAC layer, and was higher at 10 V than at 0 V. This may be related to the direct conversion of NH₄⁺-N to N₂ via electro-oxidation [35]. The average removal rate of COD_Mn in the GAC layer was only 14.81% at 10 V, indicating a drop relative to the value obtained at 0 V. Further, the average increase in UV₂₅₄ removal rate was 75.00% following an increase in voltage from 0 to 10 V. This may be a result of the generation of substances with strong oxidizing properties in the pole plate, resulting in the degradation of large-molecule organics into small organic molecules that could be easily decomposed [36]. This further led to the generation of a large quantity of substances that exhibit low removal rates and could be oxidized by potassium permanganate. The average NH₄⁺-N, COD_Mn, and UV₂₅₄ removal rates in the BC layer were 34.09%, 30.74%, and 38.46%, and 71.97%, 40.74%, and 34.83%, respectively, at 0 and 10 V, respectively. Further, at 10 V, the average removal rates of NH₄⁺-N, COD_Mn, and UV₂₅₄ in the BC layer were 12.30%, 25.93%, and 109.83% higher, respectively, than the removal rates observed for the GAC layer. The increase in the removal rate of UV₂₅₄ indicated an increase in organic material containing π-π conjugated double bonds (e.g., C=C and C=O) [37]; this may be because the rate of generation of this type of organic material was less than the rate of its degradation. The third layer (QS layer) showed the highest removal rates for all indicators. This observation indicated that UV₂₅₄ was effectively removed from the BC and QS layers, possibly owing to the role of microorganisms in biodegradation and biosorption.

Analysis and discussion based on Figure 4 showed that the first GAC biofilm electrode layer gave full play to electric field oxidation (mainly electric field oxidation), converting hard-to-degrade or large-molecule organic matter into easy-to-degrade or small-molecule organic matter. Further, the second layer, BC, gave full play to microbial degradation (mainly microbial degradation) to achieve organic matter degradation; finally, the QS layer further improved the removal of turbidity and pollutants, resulting in the efficient purification of micro-polluted water.

3.3. Analysis of the DOM Degradation Process in TL-BERs

Figure 5 shows a 3D fluorescence map (with four main fluorescence peaks: A, B, C, and D) of the feed water under a working voltage of 10 V. Fluorescence peak A (region II), Ex/Em = 220 nm/335 nm, was identified as a low-excitation wavelength tryptophan fluorescence peak, while fluorescence peak B (region III), Ex/Em = 226 nm/405 nm, was identified as a UV fulvic acid fluorescence peak. Chen [38] showed that such peaks are mainly associated with certain large molecular weights and a high degree of aromatization of polycyclic aromatic hydrocarbons (PAHs), which are difficult to degrade. Further, fluorescence peak C (region IV), Ex/Em = 276 nm/350 nm, was identified as a high-excitation wavelength tryptophan fluorescence peak, mainly corresponding to tryptophan-like components of microbial origin, while fluorescence peak D (region V), Ex/Em = 250 nm/415 nm, was identified as a humic acid fluorescence peak. This indicated that the DOM in the feed water was dominated by tryptophan-like proteins, fulvic acid, and humic acid.

Analysis of the effluent water showed substantial decreases in the intensities of all four peaks following treatment (Figure 5a). Specifically, the fluorescence intensities of peaks A, B, C, and D decreased by 36.67%, 66.22%, 67.08%, and 69.76%, respectively, indicating that the reactor was effective in removing tryptophan-like proteins, fulvic acid, and humic acid. Figure 4b shows that the area of fluorescence peak A in the GAC layer increased from 2370.1 to 3632.2 a.u., representing an increase of 53.25%. This observation indicated that tryptophan, with a lower excitation wavelength, was produced in the influent water after passing through the GAC layer. Meanwhile, the intensity of fluorescence peak C also increased from 1692.6 to 2082.2 a.u., indicating the generation of high-excitation wavelength tryptophan, consistent with the large increase in UV₂₅₄ in the GAC layer water shown in Figure 4. Fluorescence peaks B, C, and D were blue-shifted by 5, 5, and 20 nm, respectively, at the emission wavelengths. This blue-shift phenomenon [39,40] could be primarily attributed to oxidation that led to structural changes in some organic compounds in water owing to the decomposition of thick-ringed aromatic organic compounds into small molecules, hydrolysis of aromatic ring organic compounds into open rings, the disappearance of some specific functional groups, and decrease in the number of conjugated groups, indicating that the GAC layer mainly facilitated the degradation of DOM via electro-oxidation. The large increase in tryptophan-like substances in the GAC layer was advantageous because such substances are structurally simple and can easily be decomposed. This resulted in the improvement of the biochemical properties of the DOM before it reached the BC layer. The fluorescence peaks of tryptophan-like species largely disappeared after passage through the BC layer, indicating a substantial decrease in biochemical organic matter, and further suggesting that the microbial degradation of organic matter in the BC and QS layers played a major role in this regard. This is consistent with the data presented in Figure 4.

3.4. Analysis of High-Throughput Sequencing Results

The TL-BERs and BSSFs were sampled at both 10 and 0 V on filter media for high-throughput sequencing analysis of bacteria and the determination of Chao, Shannon, and Simpson indices for AHA, AHC, ATA, ATC, BH, and BT, as shown in Table S1. In the BSSFs, the results showed a greater Chao index for bacteria in the GAC layer than for those in the BC layer, indicating a higher community richness in the GAC layer. In the TL-BERs, the Chao indices of the bacteria in the BC layer, 1462.6117 for the anode and 1053.0059 for the cathode, were substantially higher than those observed in the GAC layer (1331.3793 for the anode and 1174.8447 for the cathode). Additionally, the abundance of bacteria in the vitrified layer was higher, confirming that the microorganisms in the BSSFs were mainly concentrated in the GAC layer. Further, in the TL-BERs, it was mainly concentrated in the BC layer, and the abundance of the anode community was substantially higher than that of the cathode community. This might be related to the aerobic environment generated by oxygen precipitation at the anode. The Shannon index of the bacteria indicated that the applied voltage conditions promoted the development of microbial community diversity in the TL-BERs, while the abundance of microorganisms at the anode was substantially higher than that at the cathode. Compared with the BSSFs, the TL-BERs showed enhanced microbial community richness and diversity.

From the above-mentioned findings, we infer that the TL-BERs can effectively change microbial populations and abundance, and enhance the diversity of bacterial communities between the filter layers in the reactor under the stimulation of an applied voltage. Simultaneously, the applied voltage changes the distribution of bacterial diversity in the filter media layer, resulting in greater bacterial diversity in the BC layer compared with the GAC layer, and at the anode compared with the cathode.

3.4.1. Effect of TL-BERs on Microbiome Levels

As shown in Figure 6, the dominant bacterial phyla were Proteobacteria, Firmicutes, Actinobacteriota, and Bacteroidota. Specifically, Proteobacteria, which can participate in the degradation of macromolecular organic matter in electric TL-BERs and are common dominant phyla in surface water bodies, accounted for approximately 56.99%, 62.36%, 44.27%, 14.66%, 49. 52%, and 34.26% of samples AHA, AHC, ATA, ATC, BH, and BT, respectively. Notably, some microorganisms with nitrogen and phosphorus removal functions exist in this phylum [41]. Further, Firmicutes, which are Gram-positive chemo-energetic bacteria that function as the main degraders of organic matter and can effectively degrade complex organic matter, accounted for approximately 4.26%, 25.63%, 1.74%, 39.79%, 3.70%, and 16.51% of the AHA, AHC, ATA, ATC, BH, and BT samples, respectively [42]. The relative abundances of Bacteroidota in the samples were 14.67%, 4.15%, 21.43%, 5.52%, 13.36%, and 10.44%, respectively [43]. Actinobacteriota are Gram-positive bacteria that can degrade cellulose and chitin and maintain high activity at low temperatures, while Roteobacteria, Firmicutes, Actinobacteriota, and Bacteroidota contain glucan campesterase genes [44] that show a strong ability to decompose cellulose-refractory organic matter. Chloroflexi, which are partly anaerobic, mainly degrade carbohydrates and amino acids [45]; however, competition exists between Chloroflexi and Bacteroidota for the same substrate [46]. Thus, the abundance of Chloroflexi was negatively correlated with that of Bacteroidota.

3.4.2. Effect of TL-BERs on Microbial Genus Levels

As shown in Figure 7, the dominant bacterial genera were Hydrogenophaga, Mycobacterium, Pseudonocardia, Erysipelothrix, SRB2_norank, and Sulfuritalea, with relative abundances of 6.97%, 5.87%, 5.20%, 4.43%, 4.14%, and 3.44%, respectively. Hydrogenophaga are straight or slightly curved rods with Gram-negative cells that play an important role in the degradation of PAHs during water treatment [47].

In summary, our in-depth study of bacterial communities in TL-BERs revealed that most genera play a role in degrading refractory substances under the combined effect of electric field stimulation and micro-polluted water domestication and that some genera can also metabolize the intermediate products of the electrochemical oxidation of refractory substances.

3.4.3. Correlation Analysis of Microbial Composition and Water Quality Physicochemical Characteristics

The variability in the abundance and diversity of microbial community structures is often closely related to environmental factors [48]. In this study, pH, DO, COD_Mn, NH₄⁺-N, and UV₂₅₄ were selected as environmental factors and microbial genera with relative abundances > 1% at the genus level were selected for correlation analysis (Mantel tests). Thus, the correlation between the water quality index of the filter layer discharge water and the structural composition of the microbial community was analyzed.

The distribution of bacterial abundance in the first and second principal axes alone explained 58.42% and 26.73% of the observed variance, respectively. Further, the environmental factors explained a total of 85.15% of the microbial population variation, indicating that the first and second principal axes can well reflect the environmental factors and microbial population on the surface of the filter media. From Figure 8a,b, the water physicochemical indices DO, NH₄⁺-N, and pH were positively correlated with each other but showed negative correlations with UV₂₅₄. Thus, it was evident that UV₂₅₄ had the most significant effect on bacterial community structure, while COD_Mn had a lesser effect. Further, among the dominant genera, those showing positive correlations with DO, NH₄⁺-N, and pH included Sideroxydans, A0839_norank, Vibrionimonas, Terrimonas, Beggiatoaceae, Mesorhizobium, Comamonadaceae, Rhodobacter, and Rhodanobacter. Conversely, the genera that showed positive correlations with UV₂₅₄ were Hydrogenophaga, Silanimonas, and Erysipelothrix.

The RDA results further confirmed that UV₂₅₄ had the strongest effect on bacterial community structure. Thus, the RDA results verified the correlation between the composition of the bacterial community with or without voltage stimulation, which played a key role in the high pollutant removal rates observed, and pollutant removal performance. This lays the foundation for further studies on the customization of bacterial communities in TL-BERs.

4. Conclusions

In this study, for the decentralized and small-scale treatment of micro-polluted water, we proposed the use of TL-BERs, which employ a process that is highly effective for the removal of contaminants, including DOM, dominated by tryptophan-like proteins, fulvic acid, and humic acid like substances, from water. Specifically, the highest pollutant removal rate was obtained when the water temperature, filter speed, and working voltage were 5–15 °C, 0.2 m/h, and 10 V, respectively. Additionally, NH₄⁺-N removal was mainly concentrated in the GAC layer, whereas COD_Mn and UV₂₅₄ removal mainly occurred in the BC and QS layers. Our results also indicated that at 10 V, the average removal rates of COD_Mn, NH₄⁺-N, and UV₂₅₄ increased by 36.38%, 39.63%, and 38.20%, respectively, year-on-year for effluent water relative to 0 V. The external voltage also had a substantial effect on the structure of bacterial communities in the biofilm on the surface of the filter media. Therefore, an appropriate voltage can induce changes in the dominant effects observed on each filter medium layer with respect to the removal of contaminants. In RDA, UV₂₅₄ showed the strongest negative correlation with bacterial community structure. In conclusion, the use of the TL-BERs resulted in the efficient degradation of pollutants and DOM in micro-polluted water via the synergistic action of microbial degradation and electric field oxidation. Therefore, our results provide a solution that can be applied in the future comprehensive treatment of decentralized and small-scale micro-polluted water sources, as well as informing decision-making in this regard. Because the filtration rate and residence time of this reactor are lower than those of conventional filter tanks, and the volume load of the entire reactor is smaller, equipment and operation costs must be carefully considered for larger-scale water treatment.