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Article

In Situ Utilization of Electron-Enhanced Degradation of Azo Dyes in a Constructed Wetland–Microbial Fuel Cell Coupling System

by
Xiangfeng Xie
1,2,
Jijing Hu
1,
Xian Cao
1,3,
Shuai Zhang
3,
Takashi Sakamaki
4 and
Xianning Li
1,*
1
College of Energy and Environment, Southeast University, Nanjing 210096, China
2
Jiangsu Academy of Environmental Industry and Technology Corporation, Nanjing 211800, China
3
Jiangsu Key Laboratory of Atmospheric Environment Monitoring and Pollution Control (AEMPC), Collaborative Innovation Center of Atmospheric Environment and Equipment Technology, Nanjing University of Information Science & Technology, Nanjing 210044, China
4
Department of Civil and Environmental Engineering, Graduate School of Engineering, Tohoku University, Aoba Aramaki 6-6-06, Sendai 980-8579, Japan
*
Author to whom correspondence should be addressed.
Sustainability 2024, 16(8), 3181; https://doi.org/10.3390/su16083181
Submission received: 26 February 2024 / Revised: 28 March 2024 / Accepted: 3 April 2024 / Published: 10 April 2024
(This article belongs to the Topic Removal of Hazardous Substances from Water Resources)
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Abstract

:
In this study, a constructed wetland was coupled with a microbial fuel cell to establish a coupled system known as the constructed wetland–microbial fuel cell (CW–MFC), utilized for the treatment of X-3B azo dye wastewater at varying concentrations. Experimental results indicated that the anodic region made the primary contributions to the discoloration of azo dyes and COD removal, with a contribution rate of 60.9–75.8% for COD removal and 57.8–83.0% for the effectiveness of discoloration. Additionally, the role of plants in the constructed wetland area could achieve the removal of small molecular substances and further discoloration. In comparison to open-circuit conditions, under closed-circuit conditions the CW–MFC effectively degraded X-3B azo dye wastewater. Under an external resistance of 2000 Ω, a maximum COD removal rate of 60.0% and a maximum discoloration rate of 85.8% were achieved for X-3B azo dye at a concentration of 100 mg/L. Improvements in the treatment efficiency of X-3B dye wastewater were achieved by altering the external resistance. Under an external resistance of 100 Ω and an influent concentration of X-3B of 800 mg/L, the COD removal rate reached 78.6%, and the decolorization rate reached 85.2%. At this point, the CW–MFC exhibited a maximum power density of 0.024 W/m3 and an internal resistance of 99.5 Ω. Spectral analysis and GC–MS results demonstrated the effective degradation of azo dyes within the system, indicating azo bond cleavage and the generation of numerous small molecular substances. Microbial analysis revealed the enrichment of electrogenic microorganisms under low external resistance conditions, where Geobacter and Trichococcus were dominant bacterial genera under an external resistance of 100 Ω, playing crucial roles in power generation and azo dye degradation within the system.

1. Introduction

With the rapid development of industrialization and urbanization, the discharge of large quantities of organic dye wastewater and its by-products from industrial production processes has become one of the significant challenges to water environmental quality [1,2]. Reactive Red X-3B azo dye is a type of dye widely used in industries such as textiles, printing and dyeing, and chemicals. The high-concentration discharge of it in wastewater not only threatens water quality but also poses potential hazards to aquatic ecosystems [3,4]. The aromatic ring structure and the presence of azo bonds in azo dyes endow them with a stable chemical structure, making them difficult to degrade biologically [5]. Therefore, the development of efficient and environmentally friendly technologies for dye wastewater treatment is urgently needed to achieve sustainable development of water resources [6].
Constructed wetlands (CW) and microbial fuel cells (MFC) represent innovative applications of wetland plants and microbial electrochemical technology in wastewater treatment. Constructed wetlands, by simulating the ecological processes of natural wetlands and leveraging the collective actions of plants, soil, and microorganisms, exhibit notable capabilities for organic matter removal [7,8]. Sethulekshmi et al. constructed a horizontal flow constructed wetland for treating textile dye wastewater, achieving a maximum discoloration rate of 72 ± 2% and COD removal rate of 85 ± 3% [9]. However, traditional constructed wetlands may face challenges in achieving high efficiency when dealing with complex organic wastewater [10]. In this context, the introduction of microbial fuel cells provides a novel approach to wastewater treatment. Microbial fuel cells, by utilizing microbial electron transfer between the anode and cathode, directly convert chemical energy from organic wastewater into electrical energy [11,12]. This technology not only achieves efficient wastewater degradation but also generates renewable electrical energy [13]. When applied to wastewater containing azo dyes, microbial fuel cells can efficiently degrade these organic dyes while simultaneously generating electricity [14,15,16]. Mousavian et al. constructed a dual-chamber microbial fuel cell with a graphite electrode modified by carbon nanotubes as the cathode. When the initial dye concentration was 80 mg/L, the decolorization efficiency of Reactive Blue 221 was 73.4% [17]; Chaijak et al. utilized Enterobacter sp. TS1L and textile dye wastewater to construct a dual-chamber microbial fuel cell system, achieving a decolorization rate of 72.96% for the dye wastewater [18]; Piskin et al. utilized a microbial fuel cell to treat azo dye wastewater and conducted response-surface optimization of reaction parameters. Under the optimal conditions, the removal efficiency of Sunfix Red S3B reached 91.26%, while the removal efficiency of COD was 21.61% [19]; Sun et al. studied the impact of anodic inoculants on azo dye discoloration in MFC, observing a 20.3% improvement in the decolorization of Congo Red azo dye when using textile dyeing sludge as an anodic inoculant [20].
The aim of this study is to achieve effective treatment of azo dye wastewater using sustainable technologies. Building upon prior research, we integrate shallow artificial wetlands with microbial fuel cells (CW–MFC). The role of wetland plants enhances the oxidation–reduction potential in the cathodic zone [21]. By leveraging the natural redox gradient of artificial wetlands, we establish a coupled system of artificial wetlands and microbial fuel cells [22]. This coupled system is then applied to the treatment of X-3B azo dye wastewater. By adjusting the different concentrations of X-3B and external resistance conditions, we investigate in depth the purification performance of this coupled system for azo dye wastewater under various environmental conditions. As a low-carbon, low-energy-consumption treatment technology that does not require external energy input and has the potential to generate electricity [23], the CW–MFC coupled system holds promise for wide application in real-world environments. It offers a new sustainable development pathway for addressing organic dye wastewater pollution issues.

2. Materials and Methods

2.1. Reactor Configuration

As shown in Figure 1, the primary body of the experimental setup was constructed using acrylic organic glass. It comprised a conical-shaped distributor at the bottom, a cylindrical main body, and an upper effluent weir. To simulate the actual conditions of shallow artificial wetlands, the apparatus chosen for this study has a relatively large diameter-to-height ratio. The column has an inner diameter of 30 cm, and the overall height of the reactor is 45 cm. The effective volume of the reactor was 7 L. The influent flowed in an upflow manner, sequentially passing through the bottom supporting layer, anode (activated carbon layer), intermediate gravel layer, and cathode (activated carbon layer). The gravel in the bottom supporting layer had a controlled particle size of 2–3 cm, while the gravel in the intermediate layer was controlled at 1–1.5 cm. The anode and cathode layers comprised granular activated carbon (average particle size of 4–6 mm) and stainless steel mesh (20 mesh). The granular activated carbon was washed with water, then soaked separately in 1 mol/L NaOH and HCl for 24 h to remove any potential contaminants. The inoculated sludge was obtained from a wastewater treatment plant in Jiangning, Nanjing City, China. The cathode was an air cathode, and water spinach was planted in the cathode area as part of the constructed wetland. The cathode and anode were connected via titanium wires and external resistance, while the parts in contact with the solution were encapsulated with epoxy resin to prevent short circuits. Sampling ports were installed from bottom to top at the inflow area of the anode, outflow area of the anode, inflow area of the cathode, and the overflow weir for sampling the effluent from the cathode area.

2.2. System Operation

Synthetic wastewater containing Reactive Red X-3B dye served as the influent in this experiment. Throughout the experimental period, the influent contained a glucose concentration of 500 mg/L. The nutrient and buffering solution formulations remained consistent with previous experiments [24]. The influent was delivered to the reactor’s bottom distribution area via a peristaltic pump. System current was adjusted by modifying the external resistance. The hydraulic retention time (HRT) for the entire experiment was 2 days. After reactor initiation, acclimation was carried out under an external resistance of 2000 Ω. Daily monitoring of cathode and anode potentials occurred until stability was achieved, indicating successful initiation. The first set of experiments focused on evaluating the CW–MFC system’s removal efficiency for different concentrations of X3-B. The second set delved into studying the purification efficacy of high-concentration X3-B dye wastewater under various current conditions. Finally, activated carbon without sludge inoculation was used as cathode and anode as an adsorption control experiment. After adsorption stabilization, the adsorption effect of cathodic and anodic activated carbon as well as of the matrix layer on X3-B dye wastewater was investigated.

2.3. Electrochemical Analysis

Voltage (V) and current (I) were recorded by the data acquisition module (DAM-3057 and DAM-3210, Art Technology Co., Ltd., Beijing, China) and automatically saved on a computer. At the conclusion of the experimental cycle, polarization curves were determined using the steady-state method. The coupled system was maintained in an open-circuit state for 12 h, and then the external resistance was varied using a variable resistor box within the range of 100,000 Ω to 5 Ω. After stabilizing for 30 min, the voltage across the variable resistor was measured, and the current was calculated using Ohm’s law. Polarization curves were plotted with current as the abscissa and voltage as the ordinate. The slope of the linear segment represented internal resistance. Current density and unit anode volume power density were calculated and plotted with current density on the abscissa and unit anode volume power density on the ordinate to form a power-density curve.

2.4. Other Analyses

Chemical oxygen demand (COD) was determined using the potassium dichromate method. For samples with concentrations exceeding the detection limit, further dilution was necessary for accurate measurement. X-3B concentration was measured using a UV-vis spectrophotometer (UV 9100B PC, LabTech, Los Angeles, CA, USA), detecting X-3B’s maximum absorption peak at 538 nm. By referencing a standard curve plotted within the range of 0–100 mg/L, X-3B concentrations were obtained. Influent and effluent samples, after freeze-drying, underwent Fourier-transform infrared spectroscopy (FTIR) (Nicolet iS50, ThermoFisher, Waltham, MA, USA) for analyzing changes in dye structure and functional groups. Intermediate degradation products were analyzed using gas chromatography–mass spectrometry (GC–MS) (7890A, Agilent, Santa Clara, CA, USA). From the MFC outlet, a 300 mL sample was filtered through a 0.45 μm membrane. The sample was thrice extracted with chromatography-grade CH2Cl2. The extracts were transferred from the lower layer of a separation funnel to a rotary evaporator and evaporated at 40 °C (under a vacuum of 0.098–0.001 mPa) to approximately 4 mL. The sample was continuously concentrated via nitrogen blow until reaching a volume of 2 mL, then analyzed using GC–MS. The GC conditions utilized a db5 mass spectrometry quartz capillary column (φ 0.32 mm × 25 m), high-purity He carrier gas at a flow rate of 39.5 cm/s, an inlet temperature of 280 °C, a transfer line temperature of 280 °C, and a solvent delay time of 4 min. The column temperature started at 60 °C for 0.5 min, was ramped to 235 °C at 25 °C/min, held for 2 min, increased to 250 °C at 2 °C/min, held for 5 min, and finally raised to 280 °C at 15 °C/min for 5 min. The injection volume was 1 μL. MS operated in full-scan mode, using the electron-impact ionization mode with an electron energy of 70 eV, ion source temperature of 280 °C, electron multiplier voltage of 1605 V, and scanning range from 45–600 m/z. Scanning electron microscopy (SEM) (Supra55, Zeiss, Oberkochen, Germany) was employed to observe the morphology of the biofilm attached to the granular activated carbon on the anode.
For a more in-depth analysis of the microbial community on the MFC anode, genomic DNA was extracted from microbial samples on the anode of activated carbon. The DNA concentration was measured using a TBS-380, and DNA purity was assessed using a NanoDrop2000. DNA integrity was verified through 1% agarose gel electrophoresis. The V3-V4 hypervariable regions of the bacterial 16S rRNA genes were amplified in the ABI GeneAmp® 9700 PCR Thermocycler (Applied Biosystems, Foster City, CA, USA). PCR products were then extracted from a 2% agarose gel and purified using the AxyPrep DNA Gel Extraction Kit (Axygen Biosciences, Union City, CA, USA) according to the manufacturer’s instructions. DNA quantification was performed using the Quantus™ Fluorometer (Promega, Madison, WI, USA). Libraries were constructed using the NEXTFLEX Rapid DNA-Seq Kit (Perkin Elmer, Waltham, MA, USA). Sequencing was carried out on the MiSeq PE300/Novaseq PE250 platforms (Illumina, San Diego, CA, USA).

3. Results and Discussion

3.1. Electricity Production Performance and Removal Efficiency

As shown in Figure 2, the operational cycle comprises three phases. Phase I is the initiation stage, during which the anodic electrogenic microorganisms gradually adapt to the environment. The anode potential decreases continuously, and the output voltage increases. After approximately 30 days of operation, the output voltage and current of each reactor stabilized, indicating successful startup. This suggests that shallow artificial wetlands can effectively couple with MFCs. In Phase II, the external resistance for all experimental setups is maintained at 2000 Ω. Control group A is set to an open circuit. Each experimental group is employed to treat X-3B azo dye wastewater at concentrations ranging from 100 mg/L to 800 mg/L. During Phase II, the output voltage of each experimental group ranges between 400 mV and 500 mV, with the output current remaining relatively consistent at around 0.2 mA. In Phase III, the external resistance of each experimental group is altered to investigate further removal of high-concentration X-3B azo dye under different output current conditions. It can be observed that the output current of various experimental groups remains relatively stable throughout the testing period. Group F exhibits the highest output current, with an average of 1.01 mA, while Group B has the smallest output current, averaging 0.07 mA. Finally, the impact of current intensity on the anodic microbial community is assessed at the end of the testing period.
Figure 3 displays the polarization curves and power-density curves of each device under varying external resistance conditions for the MFCs. As the external resistance of the MFCs increases, it is evident that the slope of the polarization curves for each reactor steadily rises. This suggests a positive correlation between the changes in internal resistance and external resistance for each reactor. The fitted results indicate that when the external resistances of the reactors are set to 100 Ω, 200 Ω, 700 Ω, 2000 Ω, and 6000 Ω, the corresponding internal resistances of the MFCs are 99.5 Ω, 229.4 Ω, 247.8 Ω, 838.1 Ω, and 1162 Ω, respectively. Notably, the open-circuit condition exhibits the highest internal resistance, fitted at 1177 Ω. Additionally, with the increase in MFC external resistance, there is a decreasing trend observed in the maximum power density for each reactor. The fitted results demonstrate that when the external resistances of the reactors are 100 Ω, 200 Ω, 700 Ω, 2000 Ω, and 6000 Ω, the maximum power densities of the MFCs are 0.024 W/m3, 0.023 W/m3, 0.022 W/m3, 0.007 W/m3, and 0.009 W/m3, respectively. Notably, the open-circuit condition shows the lowest maximum power density, fitted at 0.004 W/m3.
Figure 4a,b illustrate the COD removal and decolorization rates of CW–MFC when handling different concentrations of X-3B. It can be observed that, as X-3B concentrations escalate from 100 mg/L to 800 mg/L, the average COD removal rate of CW–MFC shows a declining trend. This aligns with the findings of Rathour et al., that, as dye wastewater concentration increased from 10% to 90%, COD removal by CW–MFC declined from 97.73 ± 1.61% to 70 ± 2% [25]. In this study, the maximum average COD removal rate was achieved when treating X-3B azo dye at a concentration of 100 mg/L, approximately 60.1%, which was significantly different from other experimental groups. The average COD removal rates for X-3B azo dye concentrations, ranging from 200 mg/L to 800 mg/L, varied between 49.2% and 54.1%, with no significant differences observed between groups. Additionally, when treating X-3B azo dye at a concentration of 800 mg/L, the average COD removal rate under closed-circuit conditions was 50.0%, while under open-circuit conditions it was only 35.8%, indicating a significant difference.
Similarly, the average decolorization rate of CW–MFC decreased continuously with increasing X-3B concentration. The experimental group with X-3B concentration at 100 mg/L exhibited the highest average decolorization rate, approximately 85.8%, significantly higher than other experimental groups. Conversely, the experimental group with X-3B concentration at 800 mg/L showed the lowest average decolorization rate, approximately 64.9%, while the average decolorization rate for the open-circuit control group was only 61.2%, showing significant differences from all experimental groups.
Figure 4c,d demonstrate the contribution of the adsorption effect of the cathodic and anodic region activated carbon to the decolorization of X-3B and COD removal. Under different concentration conditions, the adsorption effect exhibited similar effects on both X-3B decolorization and COD removal. The contribution of the anodic adsorption effect to COD removal and X-3B decolorization ranged from 6.5% to 7.7% and 13.1% to 16.8%, respectively. Similarly, the contribution of the cathodic adsorption effect to COD removal and X-3B decolorization ranged from 8.8% to 12.2% and 6.5% to 9.2%, respectively.
Figure 4e,f depict the COD removal and decolorization rates of CW–MFC when handling 800 mg/L X-3B under various current conditions. It can be observed that with decreasing output current in experimental groups, CW–MFC’s average COD removal rate substantially decreased. The group with an external resistance of 100 Ω exhibited the highest average COD removal and decolorization rates, reaching 78.6% and 85.2%, respectively, significantly higher than other groups. With CW–MFC external resistance varying between 200 Ω and 6000 Ω, the average COD removal rates ranged from 48.5% to 73.2%, while the decolorization rates ranged from 68.8% to 82.0%. Furthermore, under open-circuit conditions, CW–MFC exhibited the lowest average COD removal rate and decolorization rate for 800 mg/L X-3B azo dye wastewater, at 41.1% and 61.3%, respectively, significantly different from other experimental groups.
Overall, the anodic region exhibits higher COD removal and decolorization efficiency, with a contribution rate of 60.9–75.8% for COD removal and 57.8–83.0% for decolorization. As the concentration of azo dye increases, the effectiveness of pollutant removal in the anodic region decreases, consistent with the overall removal rate trend. The matrix layer under different conditions demonstrates similar adsorption capacities, with contribution rates of 5.2–10.0% for COD removal and 8.4–11.1% for decolorization. The cathodic CW area exhibited nearly consistent COD removal efficiency across different wastewater concentrations, with no noticeable pattern in its contribution to decolorization rates. Li et al. constructed a dual-chamber MFC system and separately investigated the removal efficiency of X-3B in the cathodic and anodic regions. Under optimal voltage conditions, the cathode exhibited a decolorization rate of only 0–14.16% [26]. Thung et al. provided the cathode with electron acceptors through aeration, resulting in a COD removal rate of less than 10% [27]. Compared to these studies, the cathodic region in our research demonstrated better COD removal rates (9.6–18.5%) and decolorization rates (6.1–25.2%), slightly higher than the results of adsorption experiments. The enhancement in decolorization ability of this region may be attributed to the combined adsorption of wetland plants and activated carbon particles in the CW region [28,29]. The effectiveness of COD removal may result from the combined action of activated carbon adsorption and wetland plant degradation in the cathodic region. Compared to traditional MFCs, CW–MFCs utilize artificial wetlands as the cathode, taking advantage of the aerobic environment formed by wetland plant root oxygen release without the need for cathodic aeration [30], which can expand the application range of this technology.

3.2. Intermediates and Degradation Pathways

Figure 5a,b display the FTIR spectra of the inlet and outlet of CW–MFC when treating different concentrations of X-3B azo dye wastewater, as well as the FTIR spectra of the inlet and outlet of CW–MFC under various external resistance conditions when treating X-3B azo dye wastewater. A broad absorption peak is observed around 3460 cm−1, representing the stretching vibrations of O-H and N-H [31,32]; two absorption peaks at 2336 cm−1 and 2362 cm−1 correspond to characteristic absorption peaks of CO2. The absorption peak at 1650 cm−1 corresponds to the –N=N– structure of azo dyes [33,34]. A series of sharp absorption peaks in the range of 1650 cm−1 to 1539 cm−1 represent skeletal vibrations of benzene rings, naphthalene rings, triazine structures, as well as stretching vibrations of functional groups such as C=N, C=C, and N=O [33,35]. The absorption peak near 1040 cm−1 corresponds to R-SO3- [36].
Comparing the cathode outlet with the inlet, a notable decrease in the absorption peak at 1650 cm−1 is observed in all devices, indicating the cleavage of the –N=N– of azo dyes. Correspondingly, an increase in the intensity of the skeletal vibration absorption peaks of benzene rings, naphthalene rings, and triazine structures is noticed, suggesting the formation of some smaller molecular structures after azo bond cleavage [37].
Figure 6a illustrates the UV-vis full-spectrum scanning spectra of the inlet, anode, and cathode effluents of CW–MFC treating X-3B azo dye wastewater. X-3B exhibits five distinctive absorption peaks within the UV-visible range at 233 nm, 283 nm, 330 nm, 513 nm, and 538 nm. The peak at 233 nm corresponds to the benzene ring structure [38,39], the peak at 283 nm corresponds to the triazine structure [40], the peak at 330 nm corresponds to the naphthalene ring structure [41], and the broad absorption peak at 513 nm-538 nm is associated with the conjugated N-π structure [42], related to the –N=N– structure of the X-3B azo dye and its chromophore.
It can be observed that the absorption peaks at 513 nm to 538 nm in both anode and cathode effluents are significantly lower than those in the influent, indicating the degradation of the X-3B conjugated N-π structure within the CW–MFC system, corresponding to the reason for the decolorization of the azo dye. Compared to the influent, a decrease in absorption peaks in all bands in the cathode effluent indicates the disruption of not only the conjugated structure of large molecules but also effective degradation of smaller molecules such as benzene rings and triazines in the cathodic layer. In comparison to the anode effluent, a decrease in the absorption peaks at 513 nm to 538 nm, and an increase at 233 nm and 283 nm, indicates degradation of the X-3B conjugated N-π structure in the cathodic area of the CW–MFC, generating intermediates containing benzene rings and triazine structures. However, these intermediates are not fully removed in the cathode area, leading to the presence of a certain COD concentration in the effluent, indicating that the cathode region of the CW–MFC is more effective in decolorization than in COD removal, consistent with the results shown in Figure 4a,b.
Figure 6b illustrates the UV-vis full-spectrum scan of the adsorption control experiment. The experimental results further indicate that both the cathodic and anodic activated carbon, as well as the substrate region acting as the supporting layer, can achieve slight adsorption of X-3B. Additionally, the spectra of effluents from different regions only show a decrease in intensity, with no changes in waveform, suggesting that the adsorption effect does not disrupt the molecular structure of X-3B.
Figure 6c displays the FTIR spectra of the inlet, anode effluent, and cathode effluent of the CW–MFC, further corroborating the gradual reduction in the –N=N– intensity of the X-3B azo dye after passing through both anode and cathode, and an increase in the skeletal vibration absorption peaks of benzene rings, naphthalene rings, and triazine structures after passing through the cathode, consistent with the UV results. To further identify the intermediate products of X-3B degradation, GC–MS analysis was conducted on the CW–MFC effluent at the end of different current-test periods. A total of twelve intermediate products of X-3B were detected, as shown in Table 1. These include ketones (C17H24O3, C9H14O3, C10H20O), amines (C8H17N, C7H14N2), esters (C18H26O4, C9H9NO, C11H15NO2, C20H35NO4), and other intermediates (C9H9N, C6H4C12, C16H32O3S).
Notably, only C7H14N2, containing an azo double bond, was detected among the intermediates. It is inferred that the X3-B azo double bond is primarily subjected to electron cleavage in the system, resulting in a series of intermediates containing benzene rings of large molecules. These large molecular intermediates decompose under microbial and bioelectrochemical actions, yielding smaller molecules containing benzene rings, consistent with the conclusions drawn from the FTIR results. Finally, these smaller molecules are absorbed and further mineralized by the plant roots in the CW area [43], where the benzene rings are broken, ultimately producing other simple organic compounds. The speculated degradation pathway is illustrated in Figure 7.

3.3. Microbial Community Composition

Figure 8 illustrates scanning electron microscope (SEM) images of the surface of the original granular activated carbon and the granular activated carbon retrieved from the anode region after the operational cycle. The original granular activated carbon exhibits a surface filled with pores of approximately 15 μm, 2 μm, and 0.2 μm, providing numerous attachment sites for electrogenic microorganisms [44]. After the operational cycle, evident microbial clusters are observed on the surface of the activated carbon, forming biofilm-like structures in the locations of the original larger pores. These microbial clusters exhibit abundant sticky substances secreted by bacteria, enhancing interconnection and impact resistance between microbial clusters. Additionally, this biofilm structure plays a crucial role in extracellular electron transfer, facilitating enhanced electron transfer and power generation between electrogenic microorganisms and the anode [45,46].
Figure 9a,b illustrate the differences in the anodic microbial community at the phylum and genus levels under different conditions. At the phylum level, an analysis was conducted on the top 15 relatively abundant phyla, including Proteobacteria, Patescibacteria, Firmicutes, Bacteroidota, Chloroflexi, Desulfobacterota, Actinobacteriota, Acidobacteriota, Nitrospirota, Campilobacterota, Planctomycetota, Myxococcota, Gemmatimonadota, Verrucomicrobiota, and unclassified. Compared to the original inoculum, the abundance of Proteobacteria significantly increased in both the control and experimental groups. Typically, anodic microbial communities in microbial fuel cells are primarily dominated by Proteobacteria [47], which display electrochemical activity. Furthermore, Proteobacteria are associated with the biodegradation of recalcitrant organic compounds [48], explaining the observed increase in their abundance in both the control and experimental groups. Patescibacteria, which was the dominant phylum in the original inoculum with an abundance of 36.34%, is believed to participate in denitrification processes [49], and its abundance notably decreased in both the experimental and control groups. Firmicutes are also commonly found EAB bacteria [50,51], involved in extracellular electron transfer, with abundances ranging from 11.33% to 11.59% in all experimental and control groups. An increased abundance of Chloroflexi was observed in both the experimental and control groups, and previous studies suggest that Chloroflexi can degrade recalcitrant pollutants under anaerobic conditions [52], which might explain the increase in its abundance. Desulfobacterota, known to host numerous electroactive microorganisms [53], usually highly enrich at MFC anodes. In this experiment, the abundance of Desulfobacterota increased with higher external circuit currents. In the original inoculum, the abundance of Desulfobacterota was only 0.28%. However, under CK, I-min, and I-max conditions, the abundance of Desulfobacterota escalated to 8.78%, 13.70%, and 23.57%, respectively, indicating that lower external resistance conditions favor Desulfobacterota enrichment.
At the genus level, the higher-abundance microbes in the experimental group include Geobacter, Trichococcus, Lactococcus, Nitrospira, and Citrifermentans. Geobacter, an important electrogenic microorganism, exhibited a negative correlation between its relative abundance in different reactors and the external resistance. Its abundance was 30.23%, 22.17%, and 14.46% under minimum external resistance, maximum external resistance, and CK conditions, respectively. Geobacter was hardly detected in the original water sample. According to Rago et al., the relatively higher current intensity resulting from lower external resistance is related to the extracellular electron transfer capability of Geobacter [54]. Trichococcus, belonging to the phylum Deinococcus, showed a minor variation in abundance across the experimental groups but had a slightly increased abundance compared to the original inoculum. Trichococcus possesses the capability to degrade various complex organic compounds into simpler molecules [55]; Guo et al. suggested that Trichococcus can facilitate the degradation of methyl orange in an MFC setup [56]. Lactococcus, an electroactive microorganism, is involved in fermentation and electricity generation at the MFC anode [57], aiding in the degradation of recalcitrant organic compounds [58]. Additionally, Lactococcus acts as a dye decolorizing agent [59]. Ravadelli et al. employed an anoxic–oxic membrane bioreactor (A/O-MBR) to treat wastewater containing azo dyes, and sequencing results demonstrated that Lactococcus is capable of degrading and decolorizing azo dyes [60]. Nitrospira, a nitrifying bacterium, oxidizes nitrites to nitrates, contributing to improved denitrification rates and promoting bioelectricity generation [61]; it was enriched in the experimental groups. Citrifermentans, another electrogenic microorganism, forms biofilms on carbon materials and participates in extracellular electron transfer processes [62]. Similarly to Geobacter, the abundance of Citrifermentans exhibited a negative correlation with external resistance in the setup.

3.4. Mechanism Analysis

In this study, the X-3B azo dye wastewater, after passing through the anodic region of the CW–MFC, underwent a combined effect of anaerobic conditions and bioelectrochemical processes. This environment led to the cleavage of azo bonds, facilitating significant decolorization and COD removal. High-molecular-weight dye molecules were degraded into smaller molecules such as ketones and esters due to the anaerobic conditions and bioelectrochemical interactions in the anodic zone. These smaller molecules subsequently flowed to the cathodic zone where further removal occurred through aerobic degradation, and absorption by plant roots, activated carbon particles, and the matrix. Microbial analysis indicated the enrichment of a significant number of microorganisms in the anodic region associated with electricity generation and azo dye degradation. As the external resistance of the system decreased, the increase in external circuit current further stimulated the enrichment of anodic electrogenic bacteria. Following direct and indirect extracellular electron transfer processes, electrons transferred from the intracellular space of electrogenic bacteria to the extracellular environment [63]. The additional electrons promoted the reduction degradation of the azo dye, hence enhancing the treatment efficiency of high-concentration dye wastewater by controlling the external resistance.

4. Conclusions

This study constructed a shallow-water-layer CW–MFC system, taking into account the effects of external resistance variation and pollutant concentration on the coupled system. Through pollutant concentration detection, UV-visible spectroscopy analysis, FTIR testing, and other methods, the contributions of different regions to pollutant removal were evaluated. Additionally, the final fate of X-3B within the coupled system was assessed by combining GC–MS results. The biological anode in the coupled system primarily contributes to the degradation of azo dyes, while the cathode further aids in decolorization through mechanisms such as adsorption. Microbial community analysis revealed that under low external resistance conditions, the enrichment of electroactive microorganisms such as Geobacter and Lactococcus at the anode of CW–MFC, as well as microorganisms closely related to azo dye decolorization, such as Trichococcus, is more favorable. This can effectively improve the electrochemical performance of the system and enhance the decolorization and COD removal efficiency of the coupled system for azo dye wastewater.

Author Contributions

Methodology, X.X.; investigation, J.H.; resources, S.Z. and T.S.; writing—original draft, X.X.; writing—review & editing, X.C. and X.L.; funding acquisition, X.C. and X.L. All authors have read and agreed to the published version of the manuscript.

Funding

This work was supported by the National Natural Science Foundation of China (42077108 and 42377020), Jiangsu Provincial Innovation and Entrepreneurship Doctor Program (JSSCBS20210054), Open fund by Jiangsu Key Laboratory of Atmospheric Environment Monitoring and Pollution Control (KHK 2212).

Data Availability Statement

The datasets used and/or analyzed during the current study are available from the corresponding author upon reasonable request.

Conflicts of Interest

Xiangfeng Xie was employed by the Jiangsu Academy of Environmental Industry and Technology Corp. The remaining authors declare that the research was conducted in the absence of any commercial or financial relationships that could be construed as a potential conflict of interest.

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Figure 1. Structure of CW–MFC.
Figure 1. Structure of CW–MFC.
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Figure 2. (a) Variations in output voltage of each device during the operational phase. (b) Variations in output current of each device during the operational cycle. A–F represent different experimental setups. Phase II: Device A (CK, 800 mg/L); Device B (2000 Ω, 100 mg/L); Device C (2000 Ω, 200 mg/L); Device D (2000 Ω, 400 mg/L); Device E (2000 Ω, 600 mg/L); Device F (2000 Ω, 800 mg/L). Phase III: Device A (CK, 800 mg/L); Device B (6000 Ω, 800 mg/L); Device C (2000 Ω, 800 mg/L); Device D (700 Ω, 800 mg/L); Device E (200 Ω, 800 mg/L); Device F (100 Ω, 800 mg/L).
Figure 2. (a) Variations in output voltage of each device during the operational phase. (b) Variations in output current of each device during the operational cycle. A–F represent different experimental setups. Phase II: Device A (CK, 800 mg/L); Device B (2000 Ω, 100 mg/L); Device C (2000 Ω, 200 mg/L); Device D (2000 Ω, 400 mg/L); Device E (2000 Ω, 600 mg/L); Device F (2000 Ω, 800 mg/L). Phase III: Device A (CK, 800 mg/L); Device B (6000 Ω, 800 mg/L); Device C (2000 Ω, 800 mg/L); Device D (700 Ω, 800 mg/L); Device E (200 Ω, 800 mg/L); Device F (100 Ω, 800 mg/L).
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Figure 3. (a) Polarization curves of all devices at the end of the entire operational cycle. (b) Power-density curves.
Figure 3. (a) Polarization curves of all devices at the end of the entire operational cycle. (b) Power-density curves.
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Figure 4. (a) Contribution of different zones to COD removal when treating various concentrations of X-3B with CW–MFC. (b) Contribution of different zones to azo dye decolorization when treating different concentrations of X-3B with CW–MFC. (c) Contribution of adsorption by cathodic and anodic activated carbon to COD removal when treating different concentrations of X-3B with CW–MFC. (d) Contribution of adsorption by cathodic and anodic activated carbon to azo dye decolorization when treating different concentrations of X-3B with CW–MFC. (e) Contribution of different zones to COD removal when treating 800 mg/L of X-3B under different external resistance conditions with CW–MFC. (f) Contribution of different zones to azo dye decolorization when treating 800 mg/L of X-3B under different external resistance conditions with CW–MFC. (Different letters represent significant differences between groups).
Figure 4. (a) Contribution of different zones to COD removal when treating various concentrations of X-3B with CW–MFC. (b) Contribution of different zones to azo dye decolorization when treating different concentrations of X-3B with CW–MFC. (c) Contribution of adsorption by cathodic and anodic activated carbon to COD removal when treating different concentrations of X-3B with CW–MFC. (d) Contribution of adsorption by cathodic and anodic activated carbon to azo dye decolorization when treating different concentrations of X-3B with CW–MFC. (e) Contribution of different zones to COD removal when treating 800 mg/L of X-3B under different external resistance conditions with CW–MFC. (f) Contribution of different zones to azo dye decolorization when treating 800 mg/L of X-3B under different external resistance conditions with CW–MFC. (Different letters represent significant differences between groups).
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Figure 5. (a) FTIR profiles of influent and effluent effluents treated with different concentrations of X-3B azo dyes by CW–MFC. (b) FTIR patterns of influent and effluent treated with X-3B azo dye wastewater by CW–MFC under different external resistance conditions.
Figure 5. (a) FTIR profiles of influent and effluent effluents treated with different concentrations of X-3B azo dyes by CW–MFC. (b) FTIR patterns of influent and effluent treated with X-3B azo dye wastewater by CW–MFC under different external resistance conditions.
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Figure 6. (a) CW–MFC influent, anode effluent, and cathode effluent UV-visible full-spectrum scan. (b) CW–MFC influent, anode adsorption effluent, matrix layer adsorption effluent and cathode adsorption effluent UV-visible full spectrum scan. (c) CW–MFC influent, anode effluent, and cathode effluent FTIR spectra.
Figure 6. (a) CW–MFC influent, anode effluent, and cathode effluent UV-visible full-spectrum scan. (b) CW–MFC influent, anode adsorption effluent, matrix layer adsorption effluent and cathode adsorption effluent UV-visible full spectrum scan. (c) CW–MFC influent, anode effluent, and cathode effluent FTIR spectra.
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Figure 7. X-3B proposed degradation pathway.
Figure 7. X-3B proposed degradation pathway.
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Figure 8. (a) SEM image of the original activated carbon at 3000× magnification. (b) SEM image of the original activated carbon at 10,000× magnification. (c) SEM image of the anode activated carbon at 3000× magnification. (d) SEM image of the anode activated carbon at 10,000× magnification.
Figure 8. (a) SEM image of the original activated carbon at 3000× magnification. (b) SEM image of the original activated carbon at 10,000× magnification. (c) SEM image of the anode activated carbon at 3000× magnification. (d) SEM image of the anode activated carbon at 10,000× magnification.
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Figure 9. (a) Compositional variances in the anodic microbial community at the phylum level under different external resistance conditions, and (b) compositional differences of the top 50 anodic microbial communities at the genus level under different external resistance conditions.
Figure 9. (a) Compositional variances in the anodic microbial community at the phylum level under different external resistance conditions, and (b) compositional differences of the top 50 anodic microbial communities at the genus level under different external resistance conditions.
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Table 1. X-3B degradation intermediates.
Table 1. X-3B degradation intermediates.
Serial NumberMolecular FormulaStructural Formula
1C17H24O3Sustainability 16 03181 i001
2C18H26O4Sustainability 16 03181 i002
3C9H9NOSustainability 16 03181 i003
4C11H15NO2Sustainability 16 03181 i004
5C9H14O3Sustainability 16 03181 i005
6C8H17NSustainability 16 03181 i006
7C9H9NSustainability 16 03181 i007
8C20H35NO4Sustainability 16 03181 i008
9C6H4C12Sustainability 16 03181 i009
10C16H32O3SSustainability 16 03181 i010
11C10H20OSustainability 16 03181 i011
12C7H14N2Sustainability 16 03181 i012
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Xie, X.; Hu, J.; Cao, X.; Zhang, S.; Sakamaki, T.; Li, X. In Situ Utilization of Electron-Enhanced Degradation of Azo Dyes in a Constructed Wetland–Microbial Fuel Cell Coupling System. Sustainability 2024, 16, 3181. https://doi.org/10.3390/su16083181

AMA Style

Xie X, Hu J, Cao X, Zhang S, Sakamaki T, Li X. In Situ Utilization of Electron-Enhanced Degradation of Azo Dyes in a Constructed Wetland–Microbial Fuel Cell Coupling System. Sustainability. 2024; 16(8):3181. https://doi.org/10.3390/su16083181

Chicago/Turabian Style

Xie, Xiangfeng, Jijing Hu, Xian Cao, Shuai Zhang, Takashi Sakamaki, and Xianning Li. 2024. "In Situ Utilization of Electron-Enhanced Degradation of Azo Dyes in a Constructed Wetland–Microbial Fuel Cell Coupling System" Sustainability 16, no. 8: 3181. https://doi.org/10.3390/su16083181

APA Style

Xie, X., Hu, J., Cao, X., Zhang, S., Sakamaki, T., & Li, X. (2024). In Situ Utilization of Electron-Enhanced Degradation of Azo Dyes in a Constructed Wetland–Microbial Fuel Cell Coupling System. Sustainability, 16(8), 3181. https://doi.org/10.3390/su16083181

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