Introduction
China, as a major agricultural nation with a rural population nearing 500 million, faces substantial challenges due to the substantial discharge of rural domestic sewage, totaling approximately 13.3 billion cubic meters annually [1]. About 17.6 million cubic meters of sewage are generated daily [2]. This rural domestic sewage is rich in organic matter, bacteria, minerals, suspended solids, nutrients, and other pollutants [3]. However, most rural areas in China still lack comprehensive sewage collection and treatment systems. Studies have shown that the concentrations of Chemical Oxygen Demand (COD), Total Nitrogen (TN), and Total Phosphorus (TP) in this effluent are approximately 386.4 mg/L, 50.5 mg/L, and 3.7 mg/L, respectively [3–6]. However, according to DB36/1102-2019 [7], the acceptable concentration limits of COD, TN, ammonia nitrogen (NH4+-N) and TP in the first class of water pollutant discharge standards for rural domestic sewage treatment facilities in Jiangxi Province are 60 mg/L, 20 mg/L, 8 mg/L and 1 mg/L, respectively. The traditional methods of rural domestic sewage treatment in China, which primarily consist of open dumping, direct discharge, and small-scale simple treatment, fail to meet modern environmental standards [8].
Constructed wetlands (CWs), which are artificially designed, semi-ecological wastewater treatment systems, offer several advantages including ecological friendliness, low construction costs, and ease of operation and maintenance [9]. Nowadays, CWs have been successfully applied in the treatment of domestic wastewater, agricultural wastewater, industrial wastewater, mine wastewater, landfill leachate, rainwater, polluted river water, etc. [10]. CW treatment technology has more advantages in terms of construction cost, operation management and maintenance, environmental compatibility, treatment effect, etc., compared with other technologies. At the same time, it also has certain economic benefits, is suitable for the actual situation in rural areas of China and has a good application prospect in rural areas [11]. However, despite these advantages, the practical application of CWs in rural wastewater treatment faces several challenges, such as vulnerability to the influence of climate and temperature, being easily affected by plant species, matrix saturation and clogging, non-standard design and construction, large area occupation, irrational and imperfect management, continuous production and emission of greenhouse gases (GHGs), single ecological service function, etc. These unsolved problems would certainly affect the wastewater treatment effect and shorten the service life of CWs, thus hindering the popularization and application of CWs [10–12].
To address these challenges, researchers have focused on optimizing the operational conditions of CWs, enhancing the substrate materials, and improving the design configurations to strengthen the comprehensive treatment capacity of CWs [13–15]. Operational parameters such as the system water intake method and hydraulic retention time (HRT) are crucial for pollutant removal efficiency and can be optimized to improve nitrogen and phosphorus removal. Techniques like artificial aeration and tidal flow operation can increase dissolved oxygen content, thereby enhancing the removal of NH4+-N [14]. In terms of improving the configuration of constructed wetlands, combined processes can be constructed, such as the combination of multiple constructed wetlands to build a composite constructed wetland system, constructed wetlands coupled with an electrochemical system, constructed wetlands coupled with a microbial fuel cell (CW-MFC), and so on [16]. Therefore, constructed wetland combination processes can utilize the advantages of other treatment technologies to make up for some of their deficiencies, and at the same time improve the overall purification effect of the system.
Currently, electrochemical technology, as a highly efficient technology for nitrogen and phosphorus removal, does not produce secondary pollution due to its eco-friendliness, energy requirements, selectivity, ease of operation, protection, simplicity, cost-effectiveness, and low sludge production [17–18], which opens a new way to strengthen the constructed wetland technology. Applying electrochemical technology to constructed wetland systems can increase the hydraulic load of the wetland system, reduce the area of the wetland, and attenuate the influence of climatic conditions on the nitrogen and phosphorus removal effect of the system. Meanwhile, various fillers are used in the electric field coupled constructed wetland system. However, the nitrogen and phosphorus adsorption capacities of different fillers vary significantly [19–34], which can affect the overall nitrogen and phosphorus removal efficiency of the constructed wetland. Summary of properties and TN/TP removal rates of various fillers was shown in supplementary materials Table S1. Romero et al. [35] investigated that electrochemically coupled constructed wetlands, which were fully, or partially electrically conductive materials were integrated into the CW beds to allow the flow of electrons to locally enhance oxidation and reduction reactions, reducing the footprint by increasing the pollutant removal rate. In addition, electrolytic CWs have been developed as a system that combines physical and chemical treatments by considering that low temperatures affect the activity of microorganisms and plants to remove nutrients [36], which is more reliable and effective in cold climates [37, 38]. And this could be useful to alleviate the problem that the constructed wetland is greatly affected by temperature. Electrolysis not only removed phosphorus through the process of electrocoagulation it also provided electron acceptors and donors involved in microbial metabolism, such as O2 on the anode side for nitrification or H2 on the cathode side for autotrophic denitrification, which also directed the microbial community in the bed, thus accelerating the reaction rate. The depletion of oxygen in the system limited nitrification, which would in turn affect denitrification and overall nutrient removal in a single constructed wetland system when treating high loads of nutrients [39].
Microbial fuel cell (MFC) is a device that uses the metabolic activity of microorganisms to convert organic matter into electricity. MFC technology, which are derivative of fuel cell technology [40]. For energy positive wastewater treatment or renewable energy production from biomass, MFC technology is being developed. For example, MFC [41–42], MEC [43], MDC [44], MRC [45] are derivatives of fuel cell technology for these purposes. MFC technology has demonstrated a promising potential to evolve as a sustainable approach. And CW-MFC are being developed in recent years to improve the wastewater treatment capacity of wetlands while simultaneously producing electrical power. Electroactive bacteria (e.g., Desulfovibrio desulfuricans) on the anode could oxidize the organic matter and generate electrons to compensate for the cathode in a novel shallow wetland bed-coupled closed-circuit microbial electrochemical system (WB-CMES) [39]. Meanwhile, the cathode could be enriched with more functional bacteria to have synergistic effects on the reduction of oxygen, denitrogenation, and plant growth. Xu et al. [46] investigated the construction of a multibiotic cathodic wetland-microbial fuel cell (CW-MFC) system, and the results showed that in addition to the improvement of the electrical performance, the nitrification and denitrification processes were simultaneously enhanced due to the effect of bioelectrical interactions between the power generation and the system’s nitrification (rNi) and denitrification (rDe) rates. Simultaneous correction analysis showed that the amount of nitrogen removal related to electricity was almost linearly correlated with the amount of electricity generated. The benefits of CW-MFC include high treatment efficiency, electricity generation, and recalcitrant pollutant abatement [47].
Electrochemical systems generate oxygen at the anode and hydrogen at the cathode, thereby providing the essential electron donors and acceptors required for the processes of nitrification and denitrification by microorganisms, ultimately promoting nitrification. Therefore, electrochemistry provides electrons without the need for an external carbon source, avoiding secondary pollution. And the combined process of electrochemical technology and constructed wetland technology has received more and more attention from scholars. Studies showed that the removal rate of nitrate and total phosphorus in the constructed wetland system under electrochemical enhancement could reach up [48–50], and the current intensity played an important role in the transformation of nitrogen in the electrochemical coupled constructed wetland system [48]. And the efficiency and stability of the combined system for nitrogen and phosphorus removal were better than those of the single system through the construction of the FTW-SMFCs system, and the removal rates of TN and TP by electrochemical treatment were increased [51]. While the application field of this combined technology is mainly concentrated on the treatment of tailwater of sewage treatment plants [52–54], highly ammonia nitrogen wastewater treatment [55–56] and industrial wastewater treatment [57]. And lack of studies on the effects and mechanisms of nitrogen and phosphorus removal from rural domestic sewage by this combined technology.
In order to know the effect of combined electrochemically enhanced constructed wetland system on nitrogen and phosphorus removal of rural domestic sewage, and based on the strict pollution control goal of eutrophication control in China’s river and lake basins, and further improve the treatment effect of rural domestic sewage and the effluent quality. The ordinary electrochemically enhanced constructed wetland (E-BVFCW) and three-dimensional electrochemically enhanced constructed wetland (3DE-BVFCW) were constructed based on vertical flow constructed wetland for rural domestic wastewater in this study. The effects of hydraulic retention time (HRT), current intensity, and C/N ratio on in E-BVFCW and 3DE-BVFCW systems under different operating conditions were discussed, and the best experimental conditions were obtained. By comparing the nitrogen and phosphorus removal efficiencies with those of the traditional constructed wetland (BVFCW, B2VFCW), the nitrogen and phosphorus removal effects of electrochemistry-coupled constructed wetland system for rural domestic wastewater were determined.
Materials and Methods
2.1. Water Quality Parameters Simulation
The influent used in this study was a laboratory-prepared synthetic wastewater. In accordance with the discharge characteristics of rural domestic sewage [58] and reference values for the design of rural domestic sewage treatment facilities [59], the synthetic wastewater was formulated to simulate the actual water characteristics of rural domestic sewage in Xinjian District, Nanchang, Jiangxi Province, China. The influent water of the wetland system had concentrations of approximately 120–180 mg/L COD, 30 mg/L NH4+-N, 30 mg/L NO3−-N, 6–8 mg/L TP, with a pH of 6.5–7.0. Glucose was used as the carbon source to ensure uniformity in the water, and ammonium chloride (NH4Cl), potassium nitrate (KNO3), potassium dihydrogen phosphate (KH2PO4), along with various trace elements were added. All chemicals were procured from the Chemical Management Platform at Nanchang University, and their purity was confirmed as chemically pure.
2.2 Construction and Operation of Wetland System
This experiment simulated the vertical flow constructed wetland system and the main device of the experiment was a polyvinyl chloride plastic cylinder, with a height of 60 cm, and an inner diameter of 20 cm, and the main substrate fillers was shown on the attached map (supplementary materials Table S2), the fixture diagram(Fig. 1A) and the physical figure(Fig. 1B) were shown in Fig. 1. At the bottom, 10 cm gravel particles (about 8–10 mm) were laid to prevent the matrix from blocking the water inlet and to facilitate the uniform distribution of water. The upper part of the gravel particles was mainly sponge iron particles (about 3–5 mm) with a height of 10 cm, and the upper part of the sponge iron particles was laid with gravel particles with a height of 15 cm or a mixture of granular activated carbon and gravel particles. The mixing mode of the mixed matrix was 3:1 (granular activated carbon: gravel particles) by volume. Granular activated carbon (about 1–3 mm) was coconut shell granular activated carbon. The upper part of the mixed matrix was granular activated carbon (about 1–3 mm) with a height of 10 cm. The upper part of granular activated carbon was a mixed matrix of biochar and zeolite particles (about 1–3 mm) at a height of 10 cm. The mixed matrix was mixed at a volume ratio of 1:1 between biochar and zeolite particles, and the biochar was rice straw biochar. The top layer was covered with gravel particles of 3 cm in height to prevent the loss of zeolite particles and biochar by water flow and to prevent the substrate from clogging the water outlet. In the electrochemically enhanced constructed wetland system, the sponge iron particles and stainless steel mesh (40 mesh) constituted the anode area of the system, the granular activated carbon, and stainless steel mesh constituted the cathode area of the system, the anode area and cathode area of the system were connected by copper wires. The connection between the copper wires and the stainless steel mesh was coated with a waterproof adhesive, which prevented the copper wires from being corroded. The device was powered by a stable DC power supply, and the positive and negative poles of the power supply were connected to the anode and cathode zones of the system through copper wires respectively.
After the construction of the device was completed, the plants were planted on the top layer of the system, and the configured sewage (activated sludge: sewage = 2:1) was fed into the device from the bottom inlet with a pump to cultivate the strains in the device. The activated sludge needed for microbial inoculation was taken from the aeration tank of Nanchang Honggutan sewage treatment Plant and was domesticated for one month before inoculation to the experimental system. The pH was adjusted by changing the content of NaHCO3 in water so that the pH value in the device was between 7.2 and 7.8. Filled the device with the configured sewage and closed the water inlet and outlet to keep it for 24 hr so that the sewage and the substrate filler were fully mixed. Then the water pump was used to feed the water to cultivate microorganisms continuously, the sewage flowed vertically from the top of the device to the bottom of the device, and the wastewater flowed out from the bottom outlet after passing through the cathode layer and anode layer, and the other sampling ports were closed during the process. In the process of microbial culture in the device, the effluent was discharged into the bucket to form a recirculating flow wastewater treatment system to accelerate the microbial culture rate, and the process was carried out for about 10 days, and the effluent was discharged into the wastewater bucket after 10 days. After that, the trial operation stage (45 days) was started, and the configured simulated rural domestic wastewater was continuously fed into each device with a pump, at the same time, the water samples were taken from the bottom outlet once every 2 days to determine the effluent TP, NO3−-N, NH4+-N, and TN. Finally, the stability of the system voltage and the stability of the TP, NO3−-N, NH4+-N, and TN removal rates in the effluent were used to determine whether microorganisms were cultured in the system. The main parameter settings during experimental operation were shown on the attached map (supplementary materials Table S3).
After ensuring the stability of microbial culture and system voltage, the study proceeded to investigate the effects of varying hydraulic residence times (HRT), current intensity, and C/N ratios on nitrogen and phosphorus removal within the E-BVFCW and 3DE-BVFCW systems, based on the nitrogen and phosphorus content of rural domestic sewage. The system was operated with a vertical flow inlet from the top of the system and an outlet from the bottom, and water samples were collected at the bottom outlet. The unit was operated for 15 days under each experimental condition. Four different hydraulic residence times (HRT) of 8 hr, 12 hr, 16 hr, and 24 hr were set during the experiments, corresponding to hydraulic loads of 688 L / (m2 ·d), 458 L / (m2 ·d), 344 L / (m2 ·d) and 229 L / (m2 ·d), respectively, and the other operating conditions were current intensity of 10 mA and C/N of 3. HRT was changed every 23 days to investigate its effect on each wetland system, and then each wetland system run for 15 days at each HRT condition after the stable operation of the system is confirmed. The effects of four different HRT conditions on nitrogen and phosphorus removal in E-BVFCW and 3DE-BVFCW systems were investigated. Three kinds of current intensities of 10 mA, 15 mA, and 20 mA were also set, and the other operating conditions were as follows: HRT was 24 hr and C/N was 3. The effects of three different current intensities on nitrogen and phosphorus removal in E-BVFCW and 3DE-BVFCW systems were investigated. At the same time, based on the low C/N ratios of rural domestic sewage, and comparing the effect of electrochemical devices more clearly, three kinds of influent pollutant C/N were set, which were 3, 1, and 0, respectively. The other operating conditions were as follows: HRT was 24 hr, and the current intensity was 15 mA. The effects of three kinds of influent pollutants C/N on nitrogen and phosphorus removal in E-BVFCW and 3DE-BVFCW systems were investigated.
2.3. Wetland Plants
Plants were an indispensable part of constructed wetlands, which could not only absorb organic and inorganic substances in water as nutrients for their growth but also provide a good growing environment for microorganisms through the oxygen and secretion transported by the root system. In addition, wetland plants could also improve the regional climate, beautify the local environment, and generate economic benefits and other functions [60]. Different plants planted in constructed wetlands had different effects on the purification of total phosphorus (TP) and total nitrogen (TN) by planting different plants in constructed wetlands [61]. Siyuan Song et al. [62] investigated the planting of Oenanthe javanica on a free-surface constructed wetlands at low temperatures (<10°C), which significantly enhanced the ammonia nitrogen removal (65% ~ 71%) and total nitrogen removal (41% ~ 48%). The concentration of chemical oxygen demand was dramatically increased (about 3 ~ 4 times). Zhou et al. [63] found that the changes in the concentration of Oenanthe javanica in polluted water could provide considerable economic and social benefits. Therefore, considering certain removal rates and economic benefits, Oenanthe javanica was selected as a wetland plant in this experiment with a planting density of 127 plants / m2.
2.4. Water Quality Parameters Determination
Experiments in conventional water quality measurement methods using national standard methods, specific indicators for the determination of testing, and analyzing methods and instruments are shown on the attached map (supplementary materials Table S4).
2.5. Sample Collection
2.5.1. Water sample collection
Each system adopted the way of vertical water inflow, which was fed from the top of the system and discharged from the bottom. When the effluent index had not changed and was kept for 20 days, it could be considered that the system was stable. samples were taken from the bottom outlet of each system when the system was running stably. Samples were taken every two days for a continuous two weeks to monitor the pollutant concentrations (TP, NO3−-N, NH4+-N, and TN) in the effluent of the system and to compare and analyze the effects of nitrogen and phosphorus removal in the four systems.
2.5.2. Microbial sample collection
After six months of operation, microbial samples were collected from the anode and cathode regions and the intermediate regions in the three wetland systems of E-BVFCW, 3DE-BVFCW and B2VFCW. The sampling position of the cathode region of each system is set at 5 cm from the granular activated carbon layer (hereafter referred to as the cathode layer and marked c), the anode region is set at 5 cm from the sponge iron layer (hereafter referred to as the anode layer and marked a), and the middle region is set at 8 cm from the gravel-granular activated carbon layer (hereafter referred to as the intermediate electrode layer and marked m).
Three parallel samples were taken from each sampling point. Microorganisms are separated from the substrate surface by: After sampling at each sampling point, the sample of wetland matrix was mixed and placed in a conical bottle, about 50 mL of phosphoric acid buffer (0.01 M) was added, and the oscillating solution was oscillated in a 20 kHz oscillator for 10 min. Then, the oscillating solution was filtered with a detachable 0.22-micron filter membrane, and the filter membrane sample was stored at a low temperature of −80. Finally, the filter membrane samples were sent to Shanghai Piceno Biotechnology Co., Ltd. for high-throughput sequencing.
2.6. Statistical Analysis
All samples were set in 3 parallel sets. And data were expressed as mean standard deviation. Microsoft Excel and SPSS 25.0 were mainly used to process the experimental data, and Origin 2018 software was used to plot the data graph. The removal efficiency (%) of TP, NO3−-N, NH4+-N, and TN at a different time (weeks) and for different types of constructed wetlands were calculated from the original mean values. Analysis of variance was used for the statistical analysis of significance for results. If the significant level is more than 0.05, there is no significant difference in the statistical data.
Results and Discussion
3.1. Effect of Different Operating Conditions
3.1.1. Effect of different HRT conditions
Changes in nitrogen and phosphorus removal from E-BVFCW and 3DE-BVFCW systems under different hydraulic retention time (HRT) conditions were shown in Fig. 2. The average NO3−-N removal rates of E-BVFCW and 3DE-BVFCW systems reached 96.13 ± 2.68% and 95.68 ± 2.53% at an HRT of 24 hr, which were 19.51% and 19.42% higher than those at an HRT of 8 hr respectively (p<0.05). The average TN removal rates of the E-BVFCW and 3DE-BVFCW systems reached 59.54 ± 4.99% and 68.89 ± 5.05% at an HRT of 24 hr, which were 21.14% and 21.51% higher than those at an HRT of 8 hr, respectively(p<0.05). The average removal rates of TP were 81.33 ± 2.37% and 87.93 ± 1.46% respectively, which were 6.8% and 3.51% higher than those when HRT was 8 hr (p>0.05). The average NH4+-N removal was 44.35 ± 7.20% and 56.87 ± 3.87%, which were 4.97% and 6.31% higher than those at an HRT of 8 hr (p>0.05). With the increase of hydraulic retention time (HRT), the removal efficiencies of NO3−-N and TN were significantly improved in both E-BVFCW and 3DE-BVFCW systems, but the changes in the removal efficiencies of TP and NH4+-N were not obvious, and the nitrogen and phosphorus removal effects of the 3DE-BVFCW system were significantly better than that of E-BVFCW system.
However, when the HRT was 12 hr, the average removal rates of NO3−-N in the E-BVFCW and 3DE-BVFCW systems were 80.59 ± 5.8% and 79.75 ± 5.81%, which were 3.97% and 3.49% higher than those at an HRT of 8 hr (p>0.05), and 15.54% and 15.93% lower than those at an HRT of 24 hr (p<0.05), respectively. The average removal rates of TN were 45.86 ± 2.04% and 54.57 ± 4.29%, which were 7.26% and 7.19% higher than those of 8 hr HRT (p>0.05), and 13.68% and 14.32% lower than those of 24 hr HRT (p<0.05), respectively. The average removal rates of TP were 77.91 ± 2.47% and 85.34 ± 1.93%, respectively, which were 3.38% and 0.92% higher than those of 8 hr HRT (p>0.05), and 3.42% and 2.59% lower than those when HRT was 24 hr (p>0.05), respectively. The average removal rates of NH4+-N were 41.03% and 54.62%, respectively, which were 1.65% and 4.06% higher than those at an HRT of 8 hr (p>0.05), and 3.32% and 2.25% lower than those at an HRT of 24 hr (p>0.05), respectively.
As known that nitrogen removal from wastewater was nitrification and denitrification by microorganisms. Since each microorganism has its specific growth rate, its growth is greatly dependent on the value of HRT, resulting in the change of microbial community(microbial richness, diversity, and uniformity). According to Illumina MiSeq sequencing and correlation analysis, methyl, Bdellovibrio, and Zoogloea were significantly positively correlated with nitrogen removal rate, and HRT affected the TN removal rate by changing the abundance of these genera [64]. Shorter HRT could reduce the enrichment abundance of key bacterial phyla (Proteobacteria, Bacteroidetes) and genera (Hydrogenophaga, Filimonas, Meganema) which played an important role in nitrogen and organic matter removal. In addition, shorter HRT could increase hydraulic shear stress, hindered the enrichment of carbon and nitrogen metabolism pathways, and decreased the abundance of essential genes related to glycolysis and nitrogen metabolism. While longer HRT promoted treatment processes such as assimilation, which enhanced the removal efficiency of nitrogen and phosphorus [65]. However, excessively long HRT reactors resulted in larger reactor volumes and lower biomass productivity due to lower nutrient availability. Therefore, HRT has a great effect on the nitrogen removal efficiencies of constructed wetland systems under electrochemical enhancement. While the prolonging HRT could enhance the TP removal efficiency of E-BVFCW and 3DE-BVFCW systems to a certain extent, but the enhancement effect was not obvious, which was consistent with the study results of Wei [66].
3.1.2. Effect of different current intensities conditions
Electrolysis technology is an efficient and environmentally sustainable method, which is beneficial to the reduction of nitrogen nutrients. This reduction was achieved through direct or indirect electron transfer between molecules with anodes and cathodes [67]. At the same time, coagulants could be generated in situ by electrolysis of sacrificial anodes such as Al and Fe. Therefore, the removal of phosphorus was facilitated by the formation of the Al/Fe - hydroxyl - phosphorus complexes or by the adsorption of phosphorus by Al/Fe (III) hydrolysates [68]. In addition, the abundance and activity of nitrogen-functional microorganisms were higher in the electrolysis auxiliary systems [69]. Therefore, the current intensity is also one of the factors that we must discuss.
Changes in nitrogen and phosphorus removal from E-BVFCW and 3DE-BVFCW systems under different current intensities conditions were shown in Fig. 3. Under different current intensities, the TP removal effect is as follows: When the current intensity increased from 10 mA to 15 mA, the average TP removal rate of the 3DE-BVFCW system increased from 87.93 ± 1.46% to 90.40 ± 1.46%. At this time, the maximum TP removal rate was 91.18%, and the effluent TP concentration was 0.63 mg/L. When the current intensity increased from 15 mA to 20 mA, the TP removal rate of the 3DE-BVFCW system decreased from 90.40 ± 0.38% to 88.50 ± 1.08%, a decrease of 2.27%. The change in current intensity has little effect on the TP removal efficiency of E-BVFCW, and the change range is not obvious. when the current intensity is 10 mA, 15 mA and 20 mA, the average removal rate of TP is 81.33 ± 2.37%, 80.43 ± 0.72% and 83.50 ± 1.80%, respectively. The results show that the change of current intensity has little effect on the TP removal rate of E-BVFCW and 3DE-BVFCW systems, and the changing trend of TP removal rate of the two systems is not obvious and stable.
Under the influence of different current intensities, the NO3−-N removal rate of 3DE-BVFCW system increases at first and then decreases with the increase of current intensity. When the current intensity continued to increase to 20 mA, the effluent content of NO3−-N increased significantly to 8.50 ± 2.61 mg/L, and the removal rate decreased to 74.17 ± 7.19%. The effect of current intensity on the concentration and removal rate of NO3−-N in the influent and effluent of E-BVFCW system is similar to that of 3DE-BVFCW system. When the current intensity is 10 mA, 15 mA and 20 mA, the NO3−-N removal rates of the E-BVFCW system are 93.54 ± 7.16%, 93.82 ± 0.42% and 87.54 ± 0.25%, respectively, and the effluent NO3−-N concentrations are 2.11 ± 2.23 mg/L, 1.97 ± 0.10 mg/L and 4.06 ± 0.23 mg/L, respectively. The results show that the change of current intensity has a significant effect on the NO3−-N removal rate of E-BVFCW and 3DE-BVFCW systems, and the removal efficiency increases at first and then decreases with the increase of current intensity.
For NH4+-N removal, when the current intensity is set to 10 mA, the 3DE-BVFCW system achieves a maximum NH4+-N removal rate of 63.82%, while the E-BVFCW system reaches a maximum removal rate of 55.21%. Under current intensities of 10 mA, 15 mA, and 20 mA, the effluent concentrations of NH4+-N in the 3DE-BVFCW system were recorded as 10.46 ± 0.83 mg/L, 10.24 ± 0.30 mg/L, and 10.61 ± 1.24 mg/L, respectively. Correspondingly, the average NH4+-N removal rates were 56.87 ± 3.87%, 55.08 ± 1.76%, and 54.72 ± 3.98%, respectively. The NH4+-N removal rate in the 3DE-BVFCW system appears to be minimally influenced by changes in current intensity. However, a slight decrease in removal efficiency is observed with increasing current intensity. Conversely, the NH4+-N effluent concentration in the E-BVFCW system initially decreased and then increased with varying current intensities. When the current intensity increased from 10 mA to 15 mA, the NH4+-N content in the effluent decreased from 13.50 ± 1.70 mg/L to 12.30 ± 0.21 mg/L, leading to an increase in the removal rate from 44.35 ± 7.2% to 46.45 ± 1.41%. However, at a current intensity of 20 mA, the effluent NH4+-N content increased to 14.44 ± 1.60 mg/L, resulting in a decrease in the removal rate to 38.48 ± 1.6%. These findings indicate that while the NH4+-N removal rate in the 3DE-BVFCW system is relatively stable across different current intensities, the removal rate in the E-BVFCW system significantly declines when the current intensity is too high.
For the TN removal, when the current intensity is 10 mA, 15 mA and 20 mA, the TN effluent concentration of 3DE-BVFCW is lower than that of E-BVFCW system, and the effluent concentration of 6.2 mg/L, 7.91 mg/L and 3.04 mg/L is lower than that of E-BVFCW system. The effluent concentration of TN from 10 mA to 20 mA was 20.59 ± 2.69 mg/L, 16.35 ± 0.38 mg/L and 21.46 ± 0.66 mg/L, respectively. When the current intensity increased from 10 mA to 15 mA, the TN effluent concentration of 3DE-BVFCW decreased significantly. And the removal rate of TN increased from 68.89 ± 5.05% to 76.59 ± 0.96%. When the current intensity increased from 15 mA to 20 mA, the TN effluent concentration of 3DE-BVFCW began to increase gradually, and the removal rate decreased to 70.47 ± 1.19%. In the E-BVFCW system, under three current intensities (10 mA, 15 mA and 20 mA), the average removal rates of TN were 59.54 ± 4.99%, 65.55 ± 2.12% and 66.30 ± 1.01% respectively.
As the current intensity increased, the removal rates of NO3−-N, NH4+-N, and TN in the E-BVFCW system initially increased and then decreased, with no significant change observed in the removal rate of TP, Similarly, in the 3DE-BVFCW system, the removal rates of NO3−-N, TN, and TP in 3DE-BVFCW system also increased at first and then decreased, but the change in NH4+-N removal rate was not obvious. This trend may be attributed to the fact that with the increase of current intensity, more electrons can be provided in the system to meet the needs of denitrification, and the removal rate of total ammonia in the system increases. However, if the current intensity becomes too high, nitrite reductase may be inhibited, resulting in the decrease of TN removal rate. This is corresponding to the related microbial research results: electrical stimulation can enhance bacterial metabolic activity and accelerate bacterial proliferation, and higher current intensity can increase bacterial proliferation rate [70]. However, only a limited range of current can be used to stimulate bacterial proliferation, and when the current intensity reaches 20 mA, it can lead to apoptosis [71].
When the current intensity was 15 mA, the nitrogen and phosphorus removal efficiencies of E-BVFCW and 3DE-BVFCW systems were optimal. And the nitrogen and phosphorus removal efficiencies of the 3DE-BVFCW system were significantly higher than those of the E-BVFCW system. This may be due to the fact that compared with the E-BVFCW system, the conductive granular activated carbon is added between the cathode layer and the anode layer of the 3DE-BVFCW system, which improves the efficiency of electron flow in the system, so the 3DE-BVFCW system is more obviously affected by the change of current intensity.
3.1.3. Effect of different carbon to nitrogen (C/N) ratio conditions
From 3.1.1 and 3.1.2, the optimal HRT and current intensity conditions of E-BVFCW and 3DE-BVFCW systems are obtained as 24 hr and 15 mA, respectively. Under these conditions, the effects of different influent C/N = 3, 1 and 0 on the nitrogen and phosphorus removal performance of E-BVFCW and 3DE-BVFCW systems were studied. And the effect of electrochemically enhanced constructed wetland system on nitrogen removal of wastewater with a low C/N ratio or even without a carbon source (extreme conditions) was further analyzed.
Changes in nitrogen and phosphorus removal from E-BVFCW and 3DE-BVFCW systems under different influent C/N ratio conditions were shown in Fig. 4. when the influent C/N ratio is 0 to 1 and then to 3, And the removal rates of TP in the E-BVFCW system are 86.22 ± 2.71%, 86.53 ± 2.04% and 80.43 ± 0.72% respectively. the removal rates of TP in the 3DE-BVFCW system are 89.73 ± 1.40%, 90.26 ± 1.04% and 90.40 ± 0.38% respectively. The effluent concentration and removal rate of TP in E-BVFCW and 3DE-BVFCW systems are basically not affected by the C/N ratio of influent pollutants. This may be mainly due to the fact that under the action of electrochemical enhancement, the removal of TP is mainly through the flocculation and precipitation of iron ions precipitated from the anode, and is not affected by the concentration of carbon source and TN.
For the NO3−N removal, the effluent concentration of NO3−-N in E-BVFCW and 3DE-BVFCW systems decreased and the removal rate increased with the increase of C/N ratio. When the C/N ratio increases from 1 to 3, the removal rate of NO3−-N increased from 84.92 ± 1.70% to 93.82 ± 0.42% in E-BVFCW system, and from 84.92 ± 1.70% to 93.82 ± 0.42% in 3DE-BVFCW system, respectively. When the C/N ratio of influent pollutants increases, the concentration of carbon source in water becomes higher, and NO3−-N could get enough electrons during denitrification. The C/N ratio could directly affect the denitrification efficiency of the E-BVFCW and 3DE-BVFCW systems, thus affecting the denitrification effect of the systems, so the removal rate of NO3−-N in the system is significantly improved. When the influent C/N ratio is 0, E-BVFCW and 3DE-BVFCW still have a good effect on the removal of NO3−-N, reaching about 56.28 ± 4.60% and 45.99 ± 6.86%, respectively. The results showed that the NO3−-N removal efficiency of E-BVFCW and 3DE-BVFCW systems is significantly affected by the C/N ratio, but under the action of electrochemical enhancement, a better NO3−-N removal effect could be achieved for wastewater with low carbon source or even no carbon source. This is due to the process of nitrate removal by electrochemical technology primarily involves the reduction of nitrate to N2 at the cathode, where the electrochemical system provides electrons autonomously, without the need for additional carbon sources [72,73].
For the NH4+-N removal, the removal rate of NH4+-N in E-BVFCW and 3DE-BVFCW systems increased with the increase of C/N ratio. When the C/N ratio increases from 1 to 3, the removal rate of NH4+-N increased from 23.34 ± 13.36% to 46.45 ± 1.41% in E-BVFCW system, and from 46.30 ± 6.68% to 55.08 ± 1.76% in 3DE-BVFCW system, respectively. When the C/N ratio is 0, The system still has a certain removal rate of NH4+-N, but the removal efficiency of NH4+-N was very low in E-BVFCW system and was about 29.85 ± 1.64% in 3DE-BVFCW system. Similarly, electrochemical technology primarily removes NH4+-N from sewage through electrochemical anodization, where the system provides electrons autonomously without the need for additional carbon sources [72,73].The change in C/N ratio has a significant effect on the NH4+-N removal efficiency of E-BVFCW and 3DE-BVFCW systems, and the ammonia removal efficiency of the system is significantly improved with the increase of C/N ratio.
The TN removal rate of E-BVFCW and 3DE-BVFCW systems increases with the increase of C/N ratio. When the C/N ratio increases from 1 to 3, the removal rate of TN increased from 57.18 ± 5.39% to 69.58 ± 2.43% in E-BVFCW system, and from 62.12 ± 3.28% to 78.18 ± 2.36% in 3DE-BVFCW system, respectively. When the C/N ratio is 0, E-BVFCW and 3DE-BVFCW still have a good effect on the removal of TN, reaching about 38.30 ± 4.00% and 43.39 ± 4.07%, respectively. Electrochemical technology primarily removes nitrogen from sewage through anodization, which targets NH4+-N, while cathodic reduction is used to remove inorganic nitrogen species such as nitrate and nitrite [72,73]. The electrochemical process autonomously provides electrons without the need for additional carbon sources. The higher the C/N ratio of influent pollutants, the more carbon sources can be provided for microorganisms in the system to form carbon-containing substances of microbial cells and provide energy for microbial growth, reproduction and movement, as well as more electrons for denitrification of NO3−-N. Therefore, increasing the C/N ratio can effectively improve the TN removal rate of E-BVFCW and 3DE-BVFCW systems.
3.2. Comparison of Microbial Diversity Differences and Nitrogen and Phosphorus Removal Effects in Different Systems
3.2.1. Difference of microbial diversity in the system
Under optimized operating conditions, the number of ASVs in the anode and cathode regions of the microbial communities in each system is shown in Fig. 5. Significant differences in microbial composition were observed among the E-BVFCW, 3DE-BVFCW, and B2VFCW systems, as well as between the anode and cathode regions within each system. Notably, the E-BVFCW and 3DE-BVFCW systems demonstrated a higher number of unique ASVs in both the anode and cathode regions compared to the B2VFCW system. In the intermediate electrode region, species diversity varied considerably among the three systems. The 3DE-BVFCW system had the fewest ASVs in this region, suggesting that the three-dimensional electrode configuration enhances electron transfer and conductivity, thus amplifying electrochemical effects. Meanwhile, the E-BVFCW system exhibited a significantly higher number of unique ASVs in the intermediate region compared to its anode and cathode regions. Microbial alpha diversity indices for these wetland systems are provided in supplementary materials Table S5.
Comparing the Chao1 and Observed species indices reveals that microbial richness in the anode regions of the E-BVFCW and 3DE-BVFCW systems was higher than in their corresponding cathode regions, under electrochemical influence. In the cathode region, the 3DE-BVFCW system—with electrochemical enhancement—showed slightly lower richness than the conventional B2VFCW system. However, in the anode region, the richness in 3DE-BVFCW was significantly higher than in B2VFCW, suggesting that electrochemical action substantially boosts microbial richness in the anode while slightly reducing it in the cathode.
The Shannon index further indicates that microbial diversity was greater in the cathode regions of both E-BVFCW and 3DE-BVFCW than in the anode regions. This likely reflects the aerobic conditions in the cathode region, near the water surface and plant roots, which support a more diverse microbial community due to nutrient-rich root secretions. In contrast, the Shannon index for 3DE-BVFCW was lower than that of B2VFCW, suggesting that electrochemical action can reduce overall microbial diversity.
Similarly, the Pielou index, like the Shannon index, was higher in the cathode regions of E-BVFCW and 3DE-BVFCW, indicating greater evenness in these areas. Conversely, the Pielou index for B2VFCW showed no significant difference between the anode and cathode regions, indicating uniformity. Under consistent conditions, microbial community evenness in the 3DE-BVFCW system was lower than in B2VFCW, with electrochemical effects leading to reduced richness, diversity, and evenness in the cathode region.
3.2.2. Comparison of TP removal rates in different systems
Under the same operating conditions, the effluent TP concentrations and removal rates for the BVFCW, E-BVFCW, 3DE-BVFCW, and B2VFCW systems are shown in supplementary materials Fig. S1. The average TP removal rates were 64.26 ± 5.42%, 78.92 ± 3.57%, 88.21 ± 3.03%, and 68.76 ± 4.83%, respectively. The effluent TP concentrations were 2.38 ± 0.33 mg/L, 1.42 ± 0.23 mg/L, 0.80 ± 0.23 mg/L, and 2.08 ± 0.24 mg/L, respectively. In comparison, the removal rate of PO43−-P in an integrated vertical subsurface flow constructed wetland (VSFCW) [74] was approximately 78.94%, similar to the E-BVFCW system but much lower than the 3DE-BVFCW system in this study. The TP removal rates in both E-BVFCW and 3DE-BVFCW were also significantly higher than that of a combined up-flow constructed wetland (UCW) using autoclaved aerated concrete (ACC) blocks and coconut shells as filter media, where the TP removal was only 65.5% [75]. Additionally, in a large-scale constructed wetland with long-term operation (eight years), the average TP removal rate was only 47.05 ± 20.39% [76].
Microbial community composition, structure, and diversity were analyzed for the anode, cathode, and intermediate regions in different systems (Fig. 6). At the genus level (Fig. 6A), only a small number of Tessaracoccus were found in the anode region (relative abundance: 1.30%) in the 3DE-BVFCW system, which are speculated to be fermentative phosphorus-accumulating organisms (PAOs). In contrast, the E-BVFCW system exhibited much higher Tessaracoccus abundances in both the anode (15.75%) and intermediate (7.08%) regions. Additionally, Micropruina (0.27%) was detected in the intermediate electrode region of the 3DE-BVFCW system, which was notably lower than in the E-BVFCW system (1.78%). Despite the lower abundance of phosphorus-accumulating bacteria in the 3DE-BVFCW system, it achieved a higher TP removal rate than the E-BVFCW system. This suggests that electrochemical effects, particularly phosphate flocculation and precipitation, play a more significant role in TP removal than microbial phosphorus accumulation in constructed wetlands. The TP removal rates of E-BVFCW and 3DE-BVFCW were significantly higher than those of BVFCW and B2VFCW, indicating the positive impact of electrochemistry on TP removal efficiency. Electrochemical phosphorus removal is primarily driven by electroflocculation, where metal cations (e.g., Fe2+, Al3+) released during anode electrolysis form stable precipitates with phosphate, removing it from the wastewater [37, 72, 73].
The BVFCW and B2VFCW systems also demonstrated better TP removal efficiency compared to conventional large-scale wetlands. The addition of synthetic materials like biochar, granular activated carbon, and sponge iron to these systems likely contributed to this improvement. Biochar, with its porous structure, high carbon content, and large surface area, provides significant adsorption capacity. Granular activated carbon, with its developed pore structure and abundant chemical groups, also enhances adsorption. Sponge iron, used as matrix filler in all four systems, has a large specific surface area and adsorption properties, aiding TP removal. In these systems, sponge iron acts as an electron donor, providing electrons for NO3−-N denitrification. The anode’s sponge iron releases Fe2+ under current stimulation, which is further oxidized to Fe3+, facilitating flocculation and precipitation with phosphate [77],which strengthened the electroflocculation of the system to remove phosphate from water [78]. Other studies [73] support that electrochemical phosphorus removal is achieved through co-precipitation (Fe(n+)-OH-PO4) and adsorption (FeOOH-PO4), consistent with this experiment.
The TP removal rate of the 3DE-BVFCW system was 9.29% higher than that of E-BVFCW, likely due to the addition of granular activated carbon between the cathode and anode, which provides a high surface area, good conductivity, and excellent adsorption capacity at a low cost [79]. Granular activated carbon can also regenerate itself via electrolysis within the three-dimensional electrode system [78]. Previous studies have demonstrated the effectiveness of granular activated carbon in three-dimensional electrochemical reactors (3DER) for the degradation of biuret herbicides [80] and 2,4-dichlorophenol [81]. Therefore, the addition of granular activated carbon in the 3DE-BVFCW system significantly enhanced TP removal compared to traditional constructed wetlands.
3.2.3. Comparison of NO3−-N removal rates in different systems
The changes in NO3−-N concentrations and removal rates for the effluents of the BVFCW, E-BVFCW, 3DE-BVFCW, and B2VFCW systems under identical operating conditions are shown in supplementary materials Fig. S2. The average NO3−-N removal rates for the BVFCW and B2VFCW systems were 58.62 ± 11.86% and 65.59 ± 10.18%, respectively, with average effluent concentrations of 11.68 ± 2.90 mg/L and 9.70 ± 2.63 mg/L. For the E-BVFCW system, the NO3−-N removal rate was significantly higher at 76.97 ± 7.58%, with an average effluent concentration of 6.51 ± 2.02 mg/L. Similarly, the 3DE-BVFCW system achieved a NO3−-N removal rate of 74.80 ± 7.23% and an effluent concentration of 7.08 ± 1.66 mg/L. Compared to the BVFCW and B2VFCW systems, the NO3−-N removal rates of the E-BVFCW and 3DE-BVFCW systems were significantly higher (P < 0.05), demonstrating that electrochemical enhancement can effectively improve NO3−-N removal.
The microbial community structure and relative abundance at the genus level in each region of the three systems (supplementary materials Table S6) reveal interesting patterns. In the 3DE-BVFCW system, Thauera had a relative abundance of 64.34% in the anode, 35.05% in the cathode, and 34.43% in the intermediate electrode region. The highest relative abundance in the anode suggests that aerobic denitrification is more active there. Compared to the B2VFCW system, Thauera abundance in the 3DE-BVFCW system was higher in both the anode and cathode regions, while it was lower in the intermediate region. This indicates that electrochemical action increases Thauera abundance in the anode and cathode while reducing it in the intermediate region, which correlates with the system’s improved NO3−-N removal efficiency.
Other denitrifying bacteria, such as Hydrogenophaga, Allorhizobium-Neorhizobium-Pararhizobium-Rhizobium, Pseudomonas, Denitratisoma, and Gemmatimonadetes, were also present in the 3DE-BVFCW system (Table S6). Hydrogenophaga, which can use H2 for nitrification and denitrification, had a relative abundance of 4.49%, significantly higher than in the B2VFCW system, indicating its enhancement by electrochemical action. The highest relative abundance of Hydrogenophaga in the 3DE-BVFCW system was found in the intermediate electrode region (supplementary materials Fig. S1), followed by the cathode, with the lowest abundance in the anode. This suggests that hydrogen autotrophic denitrification primarily occurs in the intermediate electrode region under electrochemical enhancement.
Denitratisoma, Allorhizobium-Neorhizobium-Pararhizobium-Rhizobium, and Pseudomonas are heterotrophic nitrifying and aerobic denitrifying bacteria, capable of denitrification under aerobic conditions. Gemmatimonadetes can reduce N2O to N2 under low-oxygen conditions [82]. Although Desulfuromonas—a Gram-negative bacterium sensitive to Fe3+ stimulation [83]—was detected in the anode and intermediate regions of the 3DE-BVFCW system, its relative abundance was low. In contrast, Desulfuromonas was more concentrated in the anode regions of the E-BVFCW and B2VFCW systems, especially in the B2VFCW system’s anode, where its relative abundance reached 6.37%.
Saccharimonadales, a typical nitrifying bacterium in glycerol-driven systems [84], was predominantly found in the cathode and intermediate regions of the 3DE-BVFCW system, with relative abundances of 4.36% and 6.05%, respectively, higher than in the E-BVFCW system. This suggests that nitrification in the 3DE-BVFCW system occurs mainly in the cathode and intermediate regions under electrochemical enhancement.
Electrochemical nitrate removal reduces NO3−-N to N2 primarily through the cathode, without the need for additional carbon sources, making the process efficient, temperature-insensitive, and easy to operate [72–73]. In the E-BVFCW system, Fe2+ precipitated from sponge iron oxidation in the anode can serve as an inorganic electron donor for NO3−-N denitrification, promoting autotrophic denitrification [85]. Electrolysis also generates H+ at the anode, which is reduced to hydrogen at the cathode, enabling hydrogen autotrophic denitrification that ultimately removes NO3−-N [86].
In comparison, other systems have shown lower NO3−-N removal rates. For example, Tan et al. [87] reported a removal rate of 54.8% in a multistage tidal stream wetland, while Chand et al. [88] achieved only 48.94% removal in a tidal system with biochar substrate. Similarly, Wang et al. [89] found that the addition of starch-FeS@PSB biochar to a vertical flow CW resulted in a removal rate of 50.77%, much lower than the results observed in this study. The enhanced effect of electrochemistry significantly improves the NO3−-N removal efficiency in the constructed wetland systems.
3.2.4. Comparison of NH4+-N removal rates in different systems
The changes in effluent NH4+-N concentrations and removal rates for each system under the same operating conditions are shown in supplementary materials Fig. S3. The average NH4+-N removal rates for the traditionally constructed BVFCW and B2VFCW wetlands were 73.02 ± 5.10% and 68.14 ± 4.89%, with corresponding effluent concentrations of 7.04 ± 1.16 mg/L and 8.39 ± 1.51 mg/L, respectively. Both systems demonstrated good NH4+-N removal efficiency, primarily through nitrification-denitrification processes [90].
The NH4+-N removal rate of the E-BVFCW system was similar to that of the BVFCW system, which used the same substrate fillers. Analysis of microbial community composition (Fig. 6B) showed a relatively high abundance of Chloroflexi in the cathode regions of both the E-BVFCW and B2VFCW systems. Chloroflexi is known for its phototrophic cooperation and ability to utilize NH4+-N and organic nitrogen for growth.
There are two main electrochemical methods for NH4+-N removal: direct oxidation and indirect oxidation. In direct oxidation, NH3 loses three electrons at the anode and is oxidized to N2. However, NH4+-N removal in the E-BVFCW system did not significantly improve compared to traditional constructed wetlands, due to the combined effects of plant uptake, matrix adsorption, and microbial conversion. In the E-BVFCW system, hydroxyl radicals produced via hydrolysis at the anode helped oxidize NH4+-N, contributing to its removal.
Several factors likely contributed to the reduced NH4+-N removal efficiency in the E-BVFCW system: 1. Oxygen Competition: Anodic electrolysis released Fe2+, which competed with nitrifying bacteria for oxygen, hindering bacterial growth and weakening nitrification [91]. 2. Anoxic Conditions: Hydrogen produced at the cathode created an anoxic environment around the cathode biofilm, further inhibiting NH4+-N nitrification [92]. 3. Endogenous Nitrate Reductase Activity: Electrochemistry may enhance endogenous nitrate reductase activity, promoting the reduction of NO3−-N and NO2−-N to NH4+-N, which could counteract NH4+-N removal.
Similarly, in the 3DE-BVFCW system, the NH4+-N removal rate was lower than that of the B2VFCW system. The 3DE-BVFCW system achieved an average NH4+-N removal rate of 63.08 ± 4.14%, with an effluent concentration of 9.68 ± 1.09 mg/L, similar to the E-BVFCW system. These results indicate that electrochemical enhancement may reduce NH4+-N removal efficiency in constructed wetland systems.
3.2.5. Comparison of TN removal rates in different systems
The changes in effluent TN concentrations and removal rates for each system under identical operating conditions are shown in supplementary materials Fig. S4. The average TN removal rates for the BVFCW, E-BVFCW, 3DE-BVFCW, and B2VFCW systems were 54.56 ± 6.57%, 58.53 ± 9.48%, 57.46 ± 8.86%, and 56.47 ± 6.90%, respectively. Corresponding effluent TN concentrations were 28.99 ± 3.94 mg/L, 26.64 ± 6.72 mg/L, 27.33 ± 6.36 mg/L, and 27.78 ± 4.25 mg/L. Among these systems, E-BVFCW had the best TN removal performance, followed closely by 3DE-BVFCW. Both electrochemically enhanced systems showed superior TN removal compared to the traditionally constructed BVFCW and B2VFCW systems. Although the E-BVFCW and 3DE-BVFCW systems had lower NH4+-N removal rates, they exhibited clear enhancements in NO3−-N removal.
At the phylum level, analysis of microbial community composition (Fig. 6B) showed that Proteobacteria and Actinobacteria dominated the communities in each region of the systems. Proteobacteria, which had the highest relative abundance (41.4% to 79.1%), plays a critical role in the nitrogen cycle due to its metabolic diversity. This phylum includes key nitrogen-removing bacteria, such as ammonia-oxidizing bacteria (AOB), nitrite-oxidizing bacteria (NOB), and denitrifying bacteria. The relative abundance of Proteobacteria was higher in the anode regions compared to the cathode regions in the E-BVFCW, 3DE-BVFCW, and B2VFCW systems. Moreover, Proteobacteria abundance was higher in both the anode and cathode regions of the 3DE-BVFCW system compared to the E-BVFCW and B2VFCW systems, suggesting that three-dimensional electrochemical enhancement promotes the growth of this phylum.
At the class level (Fig. 6C), Gammaproteobacteria and Alphaproteobacteria were dominant in all systems, playing important roles in the nitrogen cycle. The relative abundance of Gammaproteobacteria in the E-BVFCW, 3DE-BVFCW, and B2VFCW systems was as follows: E-BVFCW-c (26.6%), 3DE-BVFCW-c (54.7%), B2VFCW-c (38.61%); E-BVFCW-a (38.5%), 3DE-BVFCW-a (69.2%), B2VFCW-a (45.97%); and E-BVFCW-m (48.7%), 3DE-BVFCW-m (56.4%), B2VFCW-m (55.3%). Gammaproteobacteria was primarily found in the anode and intermediate regions, where its abundance was significantly higher than in the cathode regions. The abundance of Gammaproteobacteria in the 3DE-BVFCW system was notably higher than in the E-BVFCW and B2VFCW systems, indicating that this group is sensitive to electrochemical conditions.
At the genus level, Thauera (also known as Soxella) was the most dominant in the 3DE-BVFCW system. Thauera is a Gram-negative, rod-shaped bacterium capable of denitrification, reducing nitrite and nitrate to nitrogen gas (N2), which is then released into the atmosphere, contributing to nitrogen removal.
Alphaproteobacteria also contributed to nitrogen removal by generating electricity in the system and reducing nitrate and nitrous nitrogen in wastewater. Its relative abundance across the systems was as follows: E-BVFCW-C (13.27%), E-BVFCW-M (5.60%), E-BVFCW-A (14.81%); 3DE-BVFCW-C (9.23%), 3DE-BVFCW-M (7.86%), 3DE-BVFCW-A (9.35%); B2VFCW-C (21.17%), B2VFCW-M (14.21%), B2VFCW-A (21.31%). The abundance of Alphaproteobacteria was higher in the anode and cathode regions than in the middle electrode region. However, compared to the B2VFCW system, Alphaproteobacteria abundance was lower in both the E-BVFCW and 3DE-BVFCW systems. Additionally, the relative abundance of Deltaproteobacteria decreased under electrochemical conditions.
Actinobacteria, the second most dominant group (6.6% to 31.3%), plays a critical role in nitrogen cycling and organic matter decomposition. Actinobacteria was highly enriched in all regions of the systems, with its relative abundance showing variations due to electrochemical effects. At the class level, Actinobacteria was dominant in all systems. In the B2VFCW system, its abundance in the cathode and anode regions was 8.20% and 11.67%, respectively, while in the 3DE-BVFCW system, it was 8.6% and 10.3%. In the E-BVFCW system, Actinobacteria showed relatively high abundances of 30.8% in the cathode and 26.4% in the anode regions.
In addition to these dominant phyla, Acidobacteria, Firmicutes, Chloroflexi, and Deinococcus-Thermus also showed high relative abundance in the E-BVFCW and 3DE-BVFCW systems. Firmicutes can utilize nitrate for nitrogen removal under anaerobic conditions. Its increased abundance under electrochemical enhancement improved denitrification efficiency, enhancing nitrogen removal in these systems.
Electrochemical processes contributed to TN removal through autotrophic denitrification, with Fe2+ generated at the anode promoting NO3−-N reduction and hydrogen generated at the cathode supporting hydrogen autotrophic denitrification. As a result, both the E-BVFCW and 3DE-BVFCW systems achieved better TN removal efficiency. In contrast, traditional constructed wetlands in the Yangtze River Delta region had a TN removal rate of only 13.89 ± 54.2% [93]. Other systems using advanced materials, such as starch-FeS@PSB biochar [94] and chitosan-FeS@ peanut shell biochar [94], achieved TN removal rates of 43.71% and 52.48%, respectively, which were lower than the results obtained in the E-BVFCW and 3DE-BVFCW systems. This highlights the enhancement of TN removal efficiency through electrochemical action in constructed wetlands.
Conclusion
The most effective HRT in removing nitrogen and phosphorus from rural domestic sewage in this combined constructed wetland is 24 hr, and the optimal current intensity is 15 mA, but too high a current will inhibit the effect of nitrogen and phosphorus removal. The influent C/N ratio could significantly affect the nitrogen removal efficiency of E-BVFCW and 3DE-BVFCW systems, while the TP removal rate was unaffected by C/N ratio.
The BVFCW and B2VFCW systems can also achieve a better removal rate than ordinarily constructed wetlands by adding activated carbon particles, biochar and sponge iron. And under optimal conditions, the removal rates of TP, NO3−-N, NH4+-N, and TN in the BVFCW system could arrive 90.40 ± 0.38%, 97.81 ± 0.46%, 55.08 ± 1.76% and 76.59 ± 0.96%, respectively. The removal efficiencies of nitrogen and phosphorus from rural domestic sewage by E-BVFCW and 3DE-BVFCW systems were significantly improved under the enhancement of ordinary electrochemistry as well as three-dimensional electrochemistry. And the effluent quality can meet the relevant standards.
The coupled electrochemical-constructed wetland technology is more effective than the traditional constructed wetland system in the treatment of rural domestic sewage, and the coupled electrochemical-constructed wetland technology is applied to the treatment of rural domestic sewage. It can effectively remove nitrogen and phosphorus in the sewage and provide a certain reference basis for the subsequent rural domestic sewage treatment.
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