Advanced Wastewater Treatment: Synergistic Integration of Reverse Electrodialysis with Electrochemical Degradation Driven by Low-Grade Heat

Abstract

The reverse electrodialysis heat engine (REDHE) represents a transformative innovation that converts low-grade thermal energy into salinity gradient energy (SGE). This crucial form of energy powers reverse electrodialysis (RED) reactors, significantly changing wastewater treatment paradigms. This comprehensive review explores the forefront of this emerging field, offering a critical synthesis of key discoveries and theoretical foundations. This review begins with a summary of various oxidation degradation methods, including cathodic and anodic degradation processes, that can be integrated with RED technology. The degradation principles and characteristics of different RED wastewater treatment systems are also discussed. Then, this review examines the impact of several key operational parameters, degradation circulation modes, and multi-stage series systems on wastewater degradation performance and energy conversion efficiency in RED reactors. The analysis highlights the economic feasibility of using SGE derived from low-grade heat to power RED technology for wastewater treatment, offering the dual benefits of waste heat recovery and effective wastewater processing.

1. Introduction

The voluminous generation of organic wastewaters, a byproduct of industrial activities such as pharmaceutical synthesis, textile dyeing, and coke production, poses significant environmental challenges [1,2,3,4,5]. Characterized by refractory organic contaminants that exhibit marked resistance to biodegradation, these wastewaters often elude effective treatment through conventional biochemical methodologies. Over the past three decades, electrochemical advanced oxidation procedures (EAOPs) have gained considerable prominence, emerging as a promising solution for addressing water pollution challenges [6,7,8]. Distinguished by remarkable efficiency and efficacy, EAOPs exploit the power of in situ generated reactive species to effectively decompose a wide array of organic pollutants through oxidative processes. The technique’s inherent versatility, adaptability, and safety in ambient conditions significantly enhance its attractiveness, underscoring its potential as a leading solution in environmental remediation. Extensive scholarly inquiry has illuminated the successful application of diverse electrochemical strategies in addressing wastewaters laden with an array of contaminants [1,5]. Nonetheless, the economic viability of EAOPs has encountered barriers, primarily due to the substantial energy consumption associated with their operation [9].

Amidst the industrial production landscape, substantial volumes of organic wastewater and low-grade heat (LGH) are routinely discharged [10]. Leveraging LGH or waste heat to power enhanced electrochemical advanced oxidation processes (EAOPs) presents a compelling opportunity to dramatically curtail energy consumption and treatment costs. A promising development in this arena is the reverse electrodialysis heat engine (REDHE), a chemical heat engine that transforms LGH into salinity gradient energy (SGE) or chemical potential energy of the working fluid through thermal separation or decomposition [11,12,13,14,15,16]. This SGE is subsequently harnessed by reverse electrodialysis (RED) reactors [17,18,19], converting it into the energy required for the oxidative degradation of organic pollutants in wastewater. Adapted from relevant publications [20], this innovative technique showcases an advanced and efficient method for wastewater treatment, utilizing low-grade heat (LGH) or recovered thermal energy as its primary driving force, as depicted in Figure 1.

Transitioning from utilizing LGH to harnessing salinity gradient energy (SGE) for organic contaminant removal in wastewater marks a milestone in the development of eco-friendly methods. A critical step in this evolution is the restoration of concentration differences between high-concentration (HC) and low-concentration (LC) solutions, creating a cyclical process through selective purification of either solvent or solute. Various thermal separation strategies, including multi-effect distillation (MED) [21], membrane distillation (MD) [22], adsorption desalination (AD) [14], and air gap diffusion distillation (AGDD) [23], play key roles in this process. For instance, the implementation of a MED module within a compound apparatus by Hu et al. [24] resulted in an enhanced energetic yield of 1.01%, achieved upon configuring ten sequential evaporation stages. On the other hand, Mercer et al. [11] combined MD with reverse electrodialysis modules to recover water and energy from urinary excretions. Long et al. [25] used computational simulations to analyze the thermal and molecular transport properties of the MD-RED system, resulting in an electrical efficiency of 1.15% across temperatures from 20 °C to 60 °C. Luo et al. [26] utilized thermally labile bicarbonate-ammonium mixtures in an AD-RED system, reaching peak power densities of 0.33 W·m⁻² for the specific RED stack. Under optimal conditions, ion transport efficiency and overall energy effectiveness reached 88% and 31%, respectively. Wu et al. [23] examined the performance of the RED-AGDD integration system, noting peak system productivity of 0.014% when using a 5 M–0.20 M solution pair as the input medium.

The conversion from SGE to oxidative degradation energy is primarily facilitated by the reverse electrodialysis reactor (REDR) [27]. In the context of REDR operation, the electric potential across the membrane catalyzes redox reactions at compatible electrode surfaces [28]. Combined with different EAOPs, such as anodic oxidation [7], anodic oxidation with electrogenerated hydrogen peroxide [29], electro-Fenton (EF) [30], photoelectro-Fenton [6], and solar photoelectro-Fenton [30], RED technology can be adapted for treating various types of organic and inorganic wastewater.

In the realm of REDHE systems, thermal separation technologies, like MED, MD, AD, and AGDD, are relatively mature, enjoying widespread application in desalination and water treatment [31,32]. However, the novel treatment technology powered by SGE through REDR for organic and inorganic wastewater treatment is a burgeoning field. Researchers have developed various REDR degradation systems for different pollutants, analyzing the impact of operational parameters and system structure on degradation performance and energy conversion efficiency. However, existing reviews on RED mainly focus on advancements in RED components [33], the development of innovative hybrid systems [34], and optimization strategies for enhancing RED system performance [35]. Notably, there is still a lack of comprehensive literature reviews specifically addressing the application of RED for wastewater treatment, a critical area of increasing environmental importance. Therefore, this manuscript critically reviews the literature on utilizing SGE to sustain redox processes in REDR systems for treating wastewater contaminated with organic and inorganic pollutants. It discusses various degradation techniques applicable in REDR and examines the influence of system operation modes on degradation performance.

2. Reverse Electrodialysis Reactor

The foundational concept of using RED for electricity generation was first introduced by Pattle in 1954 [36], and since then, extensive research has been conducted on harnessing salinity gradient energy for energy production through RED technology [37,38,39]. By supplying solutions with varying salt concentrations to a RED stack, an electromotive force is generated, which can then be converted into electricity. Subsequent studies focused on advancements in RED module components, including ion-exchange membranes [40,41], spacers [42], electrodes [43], along with innovations in hybrid RED systems and the optimization of operating conditions [37], such as feed composition, flow velocity, and temperature. It was not until 2014 that Scialdone et al. [27] first demonstrated the effective use of SGE for treating various types of wastewater contaminated with organic or inorganic compounds through redox processes in a reverse electrodialysis reactor. In the following years, other research groups applied salinity gradients and REDR to treat wastewater contaminated with organics [44], chromium (VI) [45], ammonia [46] and more. It was shown that organic pollutants could be removed in the cathodic compartment through electro-Fenton processes and in the anodic compartment through direct anodic oxidation or oxidation by electrogenerated active chlorine.

In reality, the layout of a REDR closely parallels the configuration found in standard RED stacks, encompassing terminal plates, alternating sequences of anion and cation exchange membranes (AEMs and CEMs), woven separators, and electrode elements. As high-concentration and low-concentration saline solutions pass through their corresponding pathways inside the REDR, an electromotive force arises due to the counterflow migration of negative and positive ions through the AEMs and CEMs. This membrane potential drives redox reactions at both electrodes, triggering various electrochemical processes at the appropriate electrodes [44]. The specific construction is illustrated in Figure 2.

The electromotive force of a single membrane pair is derived by summing the potential differences across the positive and negative membranes, which can be calculated using the Nernst equation. Then, the electromotive force of a single membrane can be expressed as [47,48,49]:

$$E(x)={\displaystyle \frac{RT}{F}}({\alpha }_{CEM}In{\displaystyle \frac{{\gamma }_{HC}^{+}(x)C_{HC}(x)}{{\gamma }_{LC}^{+}(x)C_{LC}(x)}}+{\alpha }_{AEM}In{\displaystyle \frac{{\gamma }_{HC}^{-}(x)C_{HC}(x)}{{\gamma }_{LC}^{-}(x)C_{LC}(x)}})$$

(1)

where ${\alpha }_{AEM}$ and ${\alpha }_{CEM}$ represent the permselectivity coefficients for anion and cation exchange membranes, respectively [49]. $\gamma$ is the average ion activity coefficients. The average ion activity coefficients of anions and cations (${Na}^{+}$, ${Cl}^{-}$) in both concentrated and LC solutions are calculated using Pitzer’s virial equation [50].

The total resistance of a single membrane pair comprises the resistances from the HC and LC compartments, as well as that of IEMs [51]:

$$R_{cell}(x)=R_{HC}(x)+R_{LC}(x)+f_r(R_{CEM}+R_{AEM})$$

(2)

where $R_{AEM}$ and $R_{CEM}$ represent the resistance of the anion and cation exchange membrane, respectively. $f_r$ is used to represent the resistance correction factor to characterize the effect of salt solution concentration on membrane resistance. $R_{HC}$ and $R_{LC}$ are used to represent the solution resistance of the HC and LC solution flowing between the exchange membranes, which can be calculated using the following equation [52]:

$$R_{HC}(x)=f_y{\displaystyle \frac{d_{HC}}{{\Lambda }_{HC}(x)C_{HC}(x)}}$$

(3)

$$R_{LC}(x)=f_y{\displaystyle \frac{d_{LC}}{{\Lambda }_{LC}(x)C_{LC}(x)}}$$

(4)

where $f_y$ is the correction coefficient that takes into account the impact on the resistance due to the gasket blocking the solution [48,49], d respects the compartment’s thickness, ${\Lambda }_{HC}\left(x\right)$ and ${\Lambda }_{LC}\left(x\right)$ are the equivalent conductivities calculated from the resistance of HC and LC solutions, respectively.

The total current of a single REDR can be obtained by dividing the nodes in the X-axis direction and calculating the branch currents of each node and adding them up. The current of a single node branch can be expressed as [53]:

$$j(x)={\displaystyle \frac{N_{cell}{E}_{cell}(x)-U-E_{ele}}{N_{cell}{R}_{cell}(x)+R_{ele}(x)}}$$

(5)

where $U$ represents the terminal voltage at both ends of the external load. $R_{ele}\left(x\right)$ is the resistance of the electrode, which can be ignored [54]. Electrical potential loss ($E_{ele}$) can be described as [55]:

$$E_{ele}=E_{an}^0-E_{ca}^0+\left|{\eta }_{an}\right|+\left|{\eta }_{ca}\right|$$

(6)

where $E_{an}^0$ and $E_{ca}^0$ signify the equilibrium voltages specific to the anodic and cathodic compartments. Meanwhile, $\left|{\eta }_{an}\right|$ and $\left|{\eta }_{ca}\right|$ are the over-potential for the electrochemical reactions occurring at each electrode, which can be evaluated using the Tafel equation [55].

$$\left|\eta \right|=a_0+b_0\log j$$

(7)

where a₀ and b₀ are the empirical Tafel constants, the values of which are obtained through experimental measurement.

Consequently, the current production of each stage reactor can be evaluated as follows [39]:

$$I=b{\displaystyle {\int }_0^{L}j(x)dx}$$

(8)

where $b$ and $L$ are the width and height of the gasket compartment, respectively.

Energy efficiency is typically defined as the ratio of output energy to total input energy. Different methods are used to express the energy efficiency of REDR systems and to evaluate energy recovery from various perspectives.

The effectiveness of contaminant breakdown in a REDR is measured using the chemical oxygen demand (COD) elimination rate, which can be calculated as [56]:

$$R_{COD}=\left(1-{[COD]}_t/{[COD]}_0\right)\times 100\%$$

(9)

where [COD] is the COD concentration of wastewater.

Concurrently, the instantaneous current efficiency (ICE) is a key performance indicator in REDR systems, representing the ratio of the charge used for degrading and mineralizing organics to the total charge exchanged by the REDR system [57]. It can be expressed as

$$ICE={\displaystyle \frac{4FV}{I}}·{\displaystyle \frac{d[COD]}{dt}}$$

(10)

where F is Faraday constant, 96,485 C·mol⁻¹; V is wastewater volume in its container, m⁻³.

Energy consumption is a key metric commonly used to evaluate the efficiency of energy conversion in electrochemical oxidation and degradation processes [58]. It is specifically defined as the amount of energy required to remove a standard quantity of COD [59]. Given the dual energy consumption in the current study—where the system utilizes both SGE and electrical energy—energy consumption is divided into two categories: electricity consumption (EC) and total energy consumption (TEC). These are expressed as follows [57,60]:

$$EC={\displaystyle \frac{P_{pump}·\Delta t/3600}{32({[COD]}_0-{[COD]}_t)V}}$$

(11)

$$TEC={\displaystyle \frac{(\Delta G-P_{sys}+P_{pump})·\Delta t/3600}{32({[COD]}_0-{[COD]}_t)V}}$$

(12)

where P_pump signifies the electrical power drawn by the system’s pumping mechanisms, W; ∆t represents the duration over which treatment occurs, s; 32 is the molar mass of molecular oxygen, g·mol⁻¹; ΔG corresponds to the net amount of SGE consumed by the system during operation, W; P_sys indicates the total electrical power produced by the system, W.

The net SGE consumed by the system is quantified by the divergence between the SGE quantities at the inlet and outlet of the REDR.

$$\Delta G=SG{E}_{in}-SG{E}_{out}$$

(13)

SGE represents the Gibbs free energy disparity that arises between HC and LC solutions [61].

$$\mathrm{SGE}=2RT\left[Q_{HC}{C}_{HC}\mathrm{In}{\displaystyle \frac{C_{HC}}{C_M}}+Q_{LC}{C}_{LC}\mathrm{In}{\displaystyle \frac{C_{LC}}{C_M}}\right]$$

(14)

where R signifies the universal gas constant, 8.314 J·mol⁻¹·K⁻¹; T represents the temperature, K; C denotes the molar concentration of salt within a given solution, mol·L⁻¹. Subscript M is a solution mixed by HC and LC solutions.

3. Degradation Technology in REDR

In recent years, the field of EAOPs has seen remarkable advancements, enabling the effective degradation of a wide array of pollutants in wastewater. However, these technologies often necessitate substantial energy input and can lead to the creation of secondary pollutants. As a result, there has been growing interest among researchers in coupling RED technology with various oxidation and degradation processes to form REDR systems. This integration not only converts salinity gradient energy into chemical energy but also facilitates the breakdown of contaminants present in wastewater. Concurrently, this ingenious strategy markedly curtails the expenditure of energy conventionally required for such treatment processes (as illustrated in Figure 3).

Within REDR wastewater degradation systems, the process can be categorized into cathodic and anodic degradation, depending on the location where pollutant degradation occurs. Cathodic degradation involves reduction reactions or the utilization of strong oxidizing agents generated within the cathode chamber of the REDR, primarily encompassing cathodic reduction and the electro-Fenton process [44,62]. Conversely, anodic degradation utilizes oxidation reactions or the products generated within the anode chamber of the REDR for pollutant treatment, including anodic oxidation, electro-oxidation with active chlorine, and electrocoagulation [20,45,63].

3.1. Cathodic Pollutant Removal

3.1.1. Cathodic Reduction

Chromium compounds, ubiquitous across a spectrum of industrial sectors encompassing metal plating, pigment production for paints and textiles, leather processing, dye application, printing ink composition, and timber preservation, have sparked environmental alarms due to their intrinsic toxicity. Among these compounds, hexavalent chromium (Cr(VI)) stands out as particularly hazardous, prompting extensive research efforts aimed at its safe mitigation. A leading methodology in this domain involves the chemical and electrochemical diminution of toxic Cr(VI) species to the comparatively benign Cr(III) form—a strategy widely acknowledged for its efficacy in remedying chromium-contaminated effluent streams. Direct electrochemical reduction, facilitated on carbonaceous substrates such as carbon felt or reticulated vitreous carbon, emerges as a frontrunner in this arena. Under judiciously selected operational parameters, these platforms facilitate the almost total transmutation of Cr(VI) to Cr(III), with the cornerstone reaction occurring primarily within the confines of the cathodic compartment. This transformative process is epitomized by the equation: [62]:

$$\mathrm{Cr}\left(\mathrm{VI}\right)+3{\mathrm{e}}^{-}\to \mathrm{Cr}\left(\mathrm{III}\right)$$

(15)

Coupling direct electrochemical reduction with the cathode of the RED reactor enables electrochemical reduction processes within the cathodic chamber, facilitating the degradation of Cr(VI). In 2014, Scialdone et al. innovatively employed cathodic reduction for the detoxification of Cr(VI) to its trivalent counterpart Cr(III) within the confined environment of a REDR [62]. This innovative approach successfully removed Cr (VI) while concurrently generating electric current. As anticipated, the larger reactor exhibited faster Cr (VI) removal due to its increased electrode and membrane surface area. Specifically, within both configurations of the larger and smaller REDR reactors, the Cr(VI) concentration dipped below the quantifiable threshold of 0.01 ppm remarkably swiftly: in 6 min for the larger assembly and in 30 min for its smaller counterpart. Throughout the investigative trials, meticulous monitoring of the larger REDR unit unveiled mean current and power densities hovering around 5.9 A·m⁻² and 3.8 W·m⁻², respectively. Characteristic features of the electrode compartments included parameters such as the Cr(VI) concentration, the concentration of the supporting electrolyte, and the velocity of the electrode solutions. It was observed that conditions favorable for Cr(VI) reduction, such as higher Cr(VI) concentration, led to a decrease in cathodic potential and consequently resulted in higher power densities. Beyond Cr(VI) remediation, electrochemical reduction demonstrates efficacy in mitigating an array of contaminants, including heavy metals, inorganic anions, and halogenated organics in aquatic environments [8]. Consequently, the cathodic reduction component of the REDR approach exhibits potential for versatile application across diverse industrial wastewater treatment contexts.

3.1.2. Electro-Fenton

Electro-Fenton, as the most renowned and widely utilized electrochemical advanced oxidation process grounded in Fenton’s reaction chemistry, has seen significant advancements and broad application over the past 15 years [64,65,66]. Seamlessly incorporated into the cathode compartment of a REDR is this methodology, purposed for the proficient purification of wastewater laden with organics. The electro-Fenton process is characterized by three key components [30]:

• In situ and continuous electrogeneration of hydrogen peroxide (H₂O₂): At the carbonaceous cathode, supplied with pure oxygen or air, hydrogen peroxide is electrogenerated continuously through the reduction of oxygen (O₂), as described by Equation (16). This step is critical for the initiation of the Fenton reaction.
• Introduction of Fe²⁺ catalyst: In the solution, ferrous ions (Fe²⁺) are added to act as a catalyst, facilitating the Fenton reaction by promoting the production of hydroxyl radicals (•OH) from hydrogen peroxide.
• Cathodic reduction of Fe³⁺ to Fe²⁺: The reduction of ferric ions (Fe³⁺) to ferrous ions (Fe²⁺) at the cathode, as shown in Equation (17), ensures the continuous regeneration of the Fenton’s reagent. This crucial stage is fundamental to sustaining the efficacy of the reaction dedicated to the breakdown of organic substances.

The harmonious interaction among these stages culminates in the incessant generation of Fenton’s reagent, paving the way for the creation of exceptionally potent hydroxyl radicals (•OH). These radicals exhibit remarkable efficiency in the breakdown of organic pollutants.

$${\mathrm{O}}_2{+2\mathrm{H}}^{+}{+2\mathrm{e}}^{-}\to {\mathrm{H}}_2{\mathrm{O}}_2$$

(16)

$${\mathrm{Fe}}^{3+}+{\mathrm{e}}^{-}\to {\mathrm{Fe}}^{2+}$$

(17)

$${\mathrm{Fe}}^{2+}+{\mathrm{H}}_2{\mathrm{O}}_2+{\mathrm{H}}^{+}\to {\mathrm{Fe}}^{3+}+·{\mathrm{OH}+\mathrm{H}}_2\mathrm{O}$$

(18)

$$\mathrm{Organics}+·\mathrm{OH}\to {\mathrm{H}}_2{\mathrm{O}+\mathrm{CO}}_2$$

(19)

Within the framework of a RED reactor, employing appropriate cathode materials coupled with an electrolyte solution enriched in Fe²⁺ ions and oxygenated via air bubbling enables the initiation of the electro-Fenton (EF) process. Scialdone et al. showcased the dual potential of reverse electrodialysis processes in generating electric energy and treating Acid Orange 7 (AO7)-contaminated water through the utilization of salinity gradients [27]. The electro-Fenton (EF) process was seamlessly integrated into the REDR, wherein a carbon felt substrate served as the cathodic electrode, meticulously operated under acidic conditions characterized by a pH value of precisely 2.0. In the cathode chamber, the EF process initially achieved a rapid reduction in total organic carbon by approximately 58% within the first 3 h, yet the rate of reduction subsequently decelerated, reaching a total of 60% removal after 4.5 h. The observed deceleration could be attributed to the significant impedance caused by carboxylic acids, which are byproducts of the oxidation sequence. Throughout the experiment, the electrical current densities started at approximately 20 A·m⁻² and tapered off to around 4.3 A·m⁻² by the end, with a mean power density of about 7 W·m⁻².

From most to least influential, the operational factors significantly affecting the decolorization of AO7 dye wastewater in the cathode chamber are as follows: the pH level, wastewater flow rate, airflow velocity, and FeCl₂ concentration. Xu et al. [44] confirmed that the pH of the wastewater significantly affects decolorization efficiency at the cathode, particularly under high-current conditions. The EF reaction is optimal under acidic conditions but is hindered by both excessively high and low pH levels. This phenomenon occurs due to the use of hydrogen ions in the synthesis of hydrogen peroxide at the cathode surface, which gradually increases the pH of the wastewater in the cathodic circuit. To maintain the EF reaction conditions, a continuous introduction of diluted hydrochloric acid was necessary in the cathodic loop to regulate and stabilize the pH at around 2.0. It was also observed that decolorization efficiency initially increases but subsequently decreases with an increase in the wastewater flow rate.

D’Angelo et al. demonstrated the practical implementation of a REDR system in a scaled model facility measuring 44 × 44 cm² with 500 cell pairs. This setup utilized authentic samples of diluted brackish water alongside concentrated saline extracts sourced from evaporative ponds, achieving dual outcomes of electricity production and wastewater remediation [67]. This experimental campaign was the first of its kind to showcase the REDR system’s ability to treat wastewater in a real-world setting. An initial maximum power output of 146.6 W was recorded, and total color removal was achieved in approximately 30 min, highlighting the system’s efficiency and practicality.

3.2. Anodic Pollutant Removal

3.2.1. Anodic Oxidation

Anodic oxidation stands as the most widely employed electrochemical technique for the removal of organic pollutants from wastewater, a method that has recently gained traction for its application in decolorizing and degrading dyes within aqueous solutions [68,69,70]. This process hinges on the oxidation of pollutants within an electrolytic cell through two primary mechanisms [71]: (i) direct anodic oxidation, which often results in less efficient decontamination, and (ii) chemical reactions with species electrogenerated at the anode from water discharge, notably the physisorbed hydroxyl radical (·OH), characterized by its high standard potential (E⁰ = 2.80 V vs. SHE). The resulting oxidizing species can either lead to total or partial decontamination, contingent upon operational conditions. The behavior of electrodes in AO is delineated by two distinct cases [7]: “active” and “non-active” anodes. Active anodes, typified by Pt, IrO₂, and RuO₂, contrast with non-active anodes, represented by PbO₂, SnO₂, and BDD (boron-doped diamond). The proposed model posits that for both anode types, the initial reaction involves the oxidation of water molecules, culminating in the formation of physisorbed hydroxyl radicals (M(·OH)) on the anode surface, as described by reaction (20) [6]:

$$\mathrm{M}+{\mathrm{H}}_2\mathrm{O}\to \mathrm{M}(·{\mathrm{OH})+\mathrm{H}}^{+}{+\mathrm{e}}^{-}$$

(20)

On the surface of active anodes, a robust interaction with ·OH may occur, potentially leading to the formation of higher oxides or superoxides (MO), as shown in reaction (21), when the metal oxide anode is in a high oxidation state, surpassing the standard potential for oxygen evolution (E⁰ = 1.23 V vs. SHE). This MO/M redox couple then facilitates the oxidation of organics through reaction (22) [6]:

$$\mathrm{M}(·\mathrm{OH})\to {\mathrm{MO}+\mathrm{H}}^{+}{+\mathrm{e}}^{-}$$

(21)

$$\mathrm{Organics} +\mathrm{MO}\to \mathrm{Intermedidates}+\mathrm{M}$$

(22)

Conversely, on non-active anodes, the interaction with ·OH is sufficiently weak to allow for direct reactions between organics and M(·OH), resulting in fully oxidized end products:

$$\mathrm{Organics} +\mathrm{M}(·\mathrm{OH})\to \mathrm{M}+{\mathrm{CO}}_2+{\mathrm{H}}_2\mathrm{O}$$

(23)

Typically, when the interaction between the anode and the hydroxyl radical is less pronounced, the anode exhibits diminished reactivity towards organic compounds’ oxidation. As a result, the chemical reaction with M(·OH) proceeds more rapidly [7]. This characteristic is best exemplified by the boron-doped diamond (BDD) anode, a non-active electrode, which is proposed as the optimal choice for organic treatment via anodic oxidation (AO).

In the anodic chamber of a RED reactor, the use of sodium sulfate as the supporting electrolyte, together with non-reactive electrodes like BDD, facilitates the implementation of anodic oxidation processes. Leng et al. [72] critically evaluated the efficiency of organic contaminant removal by three different types of anodes employed within a REDR system—Ti/RuO₂–IrO₂, Ti/PbO₂, and BDD. The findings indicated that, while Ti/PbO₂ exhibited the best degradation performance at the same working fluid velocity, the system’s current efficiency was maximized when BDD was used as the anode, under the same output current. At a flow velocity of 1.0 cm·s⁻¹, the COD removal rates using Ti/RuO₂–IrO₂, Ti/PbO₂, and BDD as anodic materials were 0.51, 0.57, and 0.50, respectively. However, considering the overall current and energy efficiency, the BDD electrode outperforms other electrodes as the anode for REDR systems. With an increasing wastewater flow rate, the total current efficiency (TCE) values also increased. Upon escalation of the wastewater throughput from 200 mL·min⁻¹ to 1000 mL·min⁻¹, the TCE values for the BDD anode rose from 0.54 to 1.54. Concurrently, this increase in processing throughput led to a rise in the total energy expenditure needed to sustain continuous operations.

Wang et al. [73] explored the interplay between three varied anode compositions (Ti/IrO₂–RuO₂, Ti/PbO₂, and Ti/Ti₄O₇) on phenol degradation and net power output within the REDR. Their findings revealed that Ti/Ti₄O₇ surpassed both Ti/PbO₂ and Ti/IrO₂–RuO₂ in terms of electrochemical potential, showcasing its superiority in catalyzing redox reactions. At high currents, the COD removal efficiency of active anodes was found to be inferior to that of non-active anodes. Despite producing higher output voltage, the active anodes could not offset the increased energy consumption resulting from their lower COD removal efficiency. In contrast, Ti/PbO₂ and Ti/Ti₄O₇ anodes resulted in lower energy consumption overall within the REDR system. Under the tested conditions, the employment of Ti/Ti₄O₇ anodes notably boosted both mineralization performance and energy efficiency. Specifically, with Ti/Ti₄O₇ anodes, the REDR system exhibited a COD removal rate of 79.9% and consumed 125.0 kWh·kg COD⁻¹ at an output current of 0.2 A.

Subsequently, Leng et al. [53] developed a mathematical model for organic mineralization in REDR to predict the COD removal efficiency and total energy consumption. The simulation outcomes from this mineralization model closely correlated with the experimental data. Leng et al. [63] implemented a BDD anode paired with carbon felt within the REDR system, investigating its application in treating wastewater and utilizing AO and EF methods. Their empirical data and model simulations collectively demonstrated that increasing the working fluid velocity and the electrode rinse solution flow rate improved COD removal efficiency and instantaneous current efficiency (ICE) while reducing the total energy consumption (TEC) of the REDR system. Moreover, an increase in the initial organic matter concentration significantly reduced TEC, even when the COD removal efficiency remained nearly constant. The electricity consumption (EC) and total energy consumption (TEC) of the REDR system can reach values of 4 kWh·kg COD⁻¹ and 65 kWh·kg COD⁻¹, respectively. These values are strikingly lower than the energy requisites of customary AO wastewater decontamination techniques. Moreover, considering that SGE can be procured from residual warmth or ambient ecological sources, the REDR system possesses considerable capacity to diminish expenditures associated with wastewater management substantially. Consequently, it presents a highly promising avenue for the wastewater treatment industry.

3.2.2. Electro-Oxidation with Active Chlorine

Chlorine, along with chlorine–oxygen entities, notably including hypochlorous acid (HClO) and hypochlorite (ClO⁻), has long held a prominent position as a powerful chemical oxidizer, conventionally utilized for purifying industrial wastewater and ensuring the sterility of drinking water [74,75,76]. Advancements in electrochemical technologies have ushered in a fresh perspective on indirect electro-oxidation strategies. Herein, chlorine species activated electrochemically, produced by the direct oxidation of chloride ions at specific anodes, function to abate organic pollutants effectively. A foundational understanding in this area suggests that subjecting aqueous solutions containing chlorides to electrolysis within an integrated electrochemical reactor leads to the direct transformation of chloride ions at the anodic interface, resulting in the formation of soluble chlorine gas. This chlorine gas rapidly undergoes hydrolysis and disproportionation to form hypochlorous acid and chloride ions. Hypochlorous acid (HClO) exists in a dynamic equilibrium with hypochlorite ions (ClO⁻) in the bulk solution, governed by a pKa of 7.55, as illustrated in the following reactions [9]:

$${2\mathrm{Cl}}^{-}\to {\mathrm{Cl}}_{2\left(\mathrm{aq}\right)}{+2\mathrm{e}}^{-}$$

(24)

$${\mathrm{Cl}}_{2\left(\mathrm{aq}\right)}+{\mathrm{H}}_2\mathrm{O}\rightleftarrows {\mathrm{HClO}+\mathrm{H}}^{+}{+\mathrm{Cl}}^{-}$$

(25)

$$\mathrm{HClO}\rightleftarrows {\mathrm{H}}^{+}{+\mathrm{ClO}}^{-}$$

(26)

At pH levels below 3, Cl₂ is the dominant active chlorine species. Between pH 3 and 8, HClO becomes the primary species. Conversely, at pH values above 8, ClO⁻ dominates the chlorine speciation. Given the redox potentials of HClO (E⁰ = 1.49 V vs. SHE), Cl₂ (E⁰ = 1.36 V vs. SHE), and ClO⁻ (E⁰ = 0.89 V vs. SHE), it is evident that the oxidation of organic compounds is more rapid in acidic media compared to alkaline environments [9]. This is due to the higher reactivity and oxidizing power of HClO and Cl₂ over ClO⁻, resulting in faster mineralization of contaminants. The general pathway for the mineralization of organic dyes, as a result, is typically initiated by the direct oxidation of the dye molecules by either HClO or Cl₂. These oxidants attack the dye molecule, leading to the cleavage of the chromophore and auxochrome groups, which are responsible for the dye’s color and stability, respectively. The cleaved organic fragments then undergo further oxidation, leading to the formation of smaller, less complex molecules. The process continues until the organic compounds are fully oxidized into carbon dioxide, water, and inorganic ions. The overall process, therefore, can be summarized as follows in reaction (27). Nevertheless, there is a significant concern regarding the potential synthesis of deleterious, chlorinated organic byproducts—most notably chloroform—and chlorine–oxygen derivatives, encompassing chlorine dioxide, chlorite, and perchlorate ions concurrently [77].

$${\mathrm{Organics}+\mathrm{HClO}+\mathrm{Cl}}_2\to {\mathrm{H}}_2{\mathrm{O}+\mathrm{CO}}_2{+\mathrm{Cl}}^{-}+{\mathrm{H}}^{+}$$

(27)

In electro-oxidation processes, metals such as platinum (Pt), ruthenium dioxide (RuO₂), titanium dioxide (TiO₂), and iridium dioxide (IrO₂) stand out due to their superior electrocatalytic abilities, efficiently transforming chloride ions into active chlorine species [9]. Conversely, non-active anodes inadvertently promote reactions that lead to the formation of less desirable, non-oxidative chlorine derivatives through the over-oxidation of Cl₂ and HClO/ClO⁻.

Conclusively, within the anodic compartment of a RED reactor, utilizing NaCl as the supporting electrolyte in conjunction with active anodes leads to the formation of reactive chlorine species, thereby facilitating the degradation of organic pollutants. O. Scialdone et al. [27] successfully utilized electrogenerated active chlorine at the anode for the oxidation of organic compounds, a process that initially exhibited a slow reduction in total organic carbon (TOC) but ultimately achieved a significant TOC decrease, reaching nearly 76% after 4.5 h. Despite concerns regarding the potential formation of toxic chloro-organic compounds during active chlorine power generation, HPLC-MS analysis revealed no such compounds were produced within the REDR system. Xu et al. [44] investigated the effects of NaCl concentration and wastewater flow rate on degradation efficiencies in REDR systems, utilizing electrogenerated active chlorine. The NaCl concentration significantly influences decolorization when it is below 0.1 M. This phenomenon arises due to the diminishment in the production of active chlorine species at the anode under low-chloride-concentration scenarios, consequently curbing the indirect contaminant degradation capability of the anode. Nonetheless, once the sodium chloride concentration exceeds 0.1 M, its influence on enhancing the decolorization efficiency plateaus and becomes negligible. Within the experimental setup, it was observed that following a treatment duration of merely 20 min, the anodic chamber achieved a remarkable 99.93% decolorization efficacy for AO7 wastewater.

In addition to dye wastewater, REDR with electro-oxidation using active chlorine has been applied to treat other types of wastewater, such as ammonia and formic acid wastewater. Zhou et al. [46] discovered that during the treatment of ammonia-laden wastewaters, trace quantities of monochloramine and dichloroamine emerged as transient byproducts, suggesting that ammonia undergoes a sequential chlorination pathway before its ultimate transformation into nitrogen. In pursuit of enhancing energy recuperation whilst simultaneously curtailing the expenditure of pumping energy, the ideal flow velocities for the rinsing solution of electrodes and the saline medium were ascertained to be 250 mL·min⁻¹ and 50 mL·min⁻¹, respectively. Maximum ammonia removal and power output density using RED reached 98% and 0.06 W·m⁻², respectively. Ma et al. [78] proposed assisted-REDR for the degradation of formic acid wastewater using electro-oxidation with active chlorine, achieving a total organic carbon (TOC) reduction of approximately 55%. Additionally, assisted-REDR offers higher TOC abatement compared to REDR and requires lower cell potentials than standard electrolysis.

Based on Fermi’s equation, Leng et al. [79] developed a straightforward semi-empirical kinetic framework to elucidate the oxidation dynamics within REDR systems treating wastewater. The apparent rate constants and transition times within the model were derived directly from empirical observations. Remarkably, the computational outcomes yielded by this theoretical construct exhibited a high degree of congruence with the tangible evidence garnered from anodic decomposition trials. Jwa et al. [80] delved into assessing the enduring resilience and practical antimicrobial effectiveness of REDR setups. At peak performance settings (30 A·m⁻² current and 424 mL·min⁻¹ flow rate), onsite disinfection within the electrode solution occurred, where recirculation operation approached total pathogen elimination via generation of HClO and ClO⁻. Operating in a continuous-flow regimen, effective microbial deactivation of seawater was sustained, averaging an electrical power consumption of 0.1 ± 0.03 W (equivalent to 0.25 ± 0.07 W·m⁻²) across a testing timeframe spanning 680 h. This underscores the fouling-resistant characteristics inherent to carbon-derived catalytic elements. Additionally, REDR effluent containing active chlorine species demonstrated near-complete disinfection when mixed with aquaculture wastewater.

3.2.3. Electrocoagulation

Coagulation is a conventional physico-chemical technique used for phase separation in the detoxification of dye effluents before their release. It involves adding coagulating agents, such as Fe³⁺ or Al³⁺ ions (typically in chloride form), to precipitate the dyes. Electrochemical technology can achieve similar results through the electrocoagulation method. In electrocoagulation, an electric current facilitates the dissolution of sacrificial anodes composed of iron (Fe) or aluminum (Al) within contaminated waters, yielding metallic cations. These ions subsequently react with hydroxide ions—whose availability hinges upon the solution’s pH—to engender Fe(II), Fe(III), or Al(III) entities [81,82,83]. Acting as potent coagulants or destabilizers, these chemical species serve to nullify charges, thereby enabling the segregation of dye particulates from the wastewater matrix. Additionally, the formed aggregates can be removed through electroflotation; in this process, they attach to hydrogen gas (H₂) bubbles nucleated at the cathode, buoyantly rising to the liquid’s surface for easy separation [83].

In the application of electrocoagulation employing an iron or steel anode, Fe²⁺ ions are liberated into the wastewater through the oxidative process occurring at the anode (standard potential E⁰ = −0.44 V vs. SHE). Conversely, at the cathode, hydroxide ions and hydrogen gas are produced through the reduction reaction occurring there (E⁰ = −0.83 V vs. SHE) [4]:

$$\mathrm{Fe}\to {\mathrm{Fe}}^{2+}+2{\mathrm{e}}^{-}$$

(28)

$${2\mathrm{H}}_2{\mathrm{O}+2\mathrm{e}}^{-}\to {2\mathrm{OH}}^{-}{+\mathrm{H}}_{2\left(\mathrm{g}\right)}$$

(29)

In the oxygen-rich environment, dissolved Fe²⁺ ions undergo oxidation to precipitate as Fe(OH)₃, forming insoluble floc structures. Once generated, these Fe(OH)₃ flocs can sequester dissolved dyes through surface complexation or electrostatic attraction, effectively removing contaminants from the aqueous medium [4].

$${4\mathrm{Fe}}^{2+}{+10\mathrm{H}}_2{\mathrm{O}+\mathrm{O}}_{2\left(\mathrm{g}\right)}\to {4\mathrm{Fe}\left(\mathrm{OH}\right)}_{3\left(\mathrm{s}\right)}{+8\mathrm{H}}^{+}$$

(30)

Therefore, when Fe (or steel) or Al is used as the anode material in a RED reactor, electrocoagulation can be utilized to adsorb and remove contaminants effectively. Jia et al. [84] successfully demonstrated concurrent hydrogen generation and dye removal within a novel integrated RED and electrocoagulation (REDR-EC) framework, leveraging electrical current sourced from salinity gradients. In this experimental design, an iron electrode functions as the anode, catalyzing the production of Fe²⁺ (which then converts to Fe³⁺) to interact with hydroxide ions (OH⁻), created at the cathode, facilitating dye contaminant elimination. Concurrently, hydrogen gas evolved at the cathode is harvested utilizing a dedicated hydrogen recovery apparatus. The results showed that, under a current of 0.4 A, the dye degradation efficiency and hydrogen output peaked at 98.3% and 150 m³·h⁻¹ under alkaline (pH = 11) and acidic (pH = 3) conditions, respectively. Furthermore, Jia et al. [45] conducted comprehensive studies to evaluate the impact of electrode characteristics—including surface area, composition, and architecture—on the efficacy of methyl orange (MO) decolorization alongside the hydrogen yield. It was observed that when an iron electrode served as the anode, it outperformed its aluminum counterpart significantly in terms of MO removal efficiency; however, the opposite held true regarding the hydrogen production capacity. Regarding cathode selection, a titanium mesh configuration proved optimal for achieving the highest MO degradation rates, while a solid titanium plate performed better in enhancing hydrogen production yields.

Zhang et al. [85] innovatively combined reverse electrodialysis with electrocoagulation, achieving simultaneous electricity generation and Cr(VI) removal for the first time. Utilizing salinity gradient energy, the system dissolved an Fe anode to produce Fe²⁺, reducing Cr(VI) and enabling its removal via coagulation as Fe³⁺. They reported 99% Cr(VI) removal within 2 h, with a peak power output density of 0.482 W·m⁻² when the salinity gradient ratio of high- to low-concentration solutions was 3/1 and 300/1, respectively.

3.3. Other Methods

In addition to the aforementioned coupling of REDR with primary electrochemical oxidation processes, other researchers have also explored integrating RED technology with novel wastewater treatment methods. Li et al. [86] introduced a groundbreaking method employing a microbial reverse-electrodialysis electrolysis cell coupled with the Fenton process to efficiently treat azo dye-contaminated wastewater. Specifically, Orange G at 400 mg·L⁻¹ was rapidly degraded with a rate constant of 1.15 ± 0.06 h⁻¹, achieving near-complete mineralization with a total organic carbon (TOC) removal of 99.6% at pH 2 while consuming only 25.93 kWh·kg TOC⁻¹. Tian et al. [87] conceived a solar-salinity energy integration system by merging a photocatalytic fuel cell with reverse electrodialysis (PRC). This innovative approach not only converted solar and salinity gradient energies into electrical power but also amplified the breakdown of Rhodamine B and boosted H₂O₂ synthesis. The bias voltage supplied by the REDR component facilitated the enhanced separation of photogenerated charge carriers on the 3D TiO₂ array photoanode, thereby improving both pollutant degradation and H₂O₂ yield. Under optimized settings, the PRC delivered a peak power density of 1500 mW·m⁻² and an energy conversion efficiency of 4.21%. Although its electricity generation performance surpassed that of a traditional photocatalytic electrolytic cell powered by desalination, it was inferior to the microbial reverse-electrodialysis electrolysis cell. Later, Tian et al. [88] refined the PRC further by incorporating a hydrogen evolution cathode, aiming to concurrently enhance electricity generation and hydrogen production while degrading stubborn organic pollutants. By capitalizing on the REDR-generated bias voltage, the team observed an increase in ampicillin removal from 58.5% to 74.2%, along with activation of the hydrogen evolution reaction, culminating in a cumulative hydrogen yield of 500 μmol.

4. Operational Mode

In REDR systems coupled with different oxidation degradation technologies, not only do the inherent properties of the wastewater affect degradation efficiency and energy consumption, but the operational mode of the REDR also plays a crucial role in degradation performance and energy conversion efficiency. Based on current research, key operational factors that influence wastewater degradation and system energy conversion performance include the output current (current density) [20], degradation circulation mode [44], and the use of multi-stage REDR-series systems [60].

4.1. Current Density

Current density (j) is a key parameter in electrochemical oxidation degradation methods, as it governs the production of oxidizing species. For example, in anodic degradation methods, j controls the amount of M(•OH) generated via Equation (20) as well as the production of indirect oxidizing agents, such as active chlorine species from Equation (24). In the EF process in REDR, j also controls the amount of electrogenerated H₂O₂ via Equation (16). Additionally, it regulates the regeneration of Fe³⁺ to Fe²⁺ via Equation (17), which, together with H₂O₂ electro-generation, dictates the amount of •OH produced from Fenton’s Equation (18). Generally, the rate of pollutant degradation increases with rising j across all electrochemical degradation technologies, as more oxidizing species are generated over time. Therefore, increasing the output current in REDR degradation systems can significantly enhance wastewater degradation efficiency. For instance, in the RED-EC system, increasing the output current from 0.15 A to 0.4 A over 120 min boosts azo dye removal from 48.59% to 79.12% [45]. This occurs as higher currents accelerate iron electrode oxidation, producing more iron ions that form hydroxyl polymers, enhancing dye molecule adsorption.

Within the REDR framework, significant increases in current density can be achieved by increasing the number of membrane pairs, accelerating the flow rate of HC and LC solution streams, amplifying the salinity gradient, or reducing REDR resistance [37,89,90]. In the RED-EC system [45], the higher current density achieved with more membrane pairs accelerated Cr(VI) removal. Furthermore, escalating the count of membrane pairs engenders an elevation in power density output, concurrent with surges in current density and cell potential. This phenomenon arises since the detrimental effects of energy dissipation from redox reactions become proportionally less significant relative to the gross power output. Particularly in configurations boasting an extensive array of cells, the decrement in potential attributed to redox processes dwindles to near insignificance juxtaposed against the aggregate cell potential [63].

The mean output current witnessed an upsurge concomitant with the escalation in the flow velocity of the saline solutions, due to the heightened electromotive force and reduced resistance, which contributed to more efficient oxidation [46]. In Scialdone et al.’s study [62], higher flow rates (400 mL·min⁻¹, 2.5 cm·s⁻¹) precipitated elevated current densities, which, in turn, expedited the elimination process of Cr(VI). The research clearly shows that the COD concentration decreases more rapidly with increasing flow velocity [72]. At flow velocities of 0.2 cm·s⁻¹, 0.6 cm·s⁻¹, and 1.0 cm·s⁻¹, the COD removal rate of MO wastewater increased to 26%, 40%, and 51% after one hour of treatment. An augmented current density catalyzes the synthesis of a greater volume of oxidizing agents at the REDR electrode surfaces; these agents subsequently serve to oxidize and decompose contaminants within the wastewater. Although a higher saline solution flow rate helped generate more power and accelerated pollutant removal, it also required more pumping energy and caused greater pressure losses [63]. As the velocity of saline solutions escalated, the energy input rose, while the maximum energy recovery decreased. An enhanced salinity disparity triggered a marked elevation, not only in current density but also in cell potential. Investigations demonstrated that the baseline power density leaped from roughly 13 W·m⁻² up to 48 W·m⁻² across scenarios characterized by diminished (NaCl 0.5 M and 0.01 M in HC and LC compartments) and magnified (NaCl 5 M and 0.01 M in HC and LC compartments) salinity gradients [91]. Moreover, attributable to the intensified output current, the RED mechanism under a pronounced salinity gradient exhibited an expedited ammonia oxidation frequency at the two-hour juncture [46].

Conversely, the operational current is also modulated by the REDR’s resistance, encompassing both its external and internal components. Ma et al. [92] systematically explored a spectrum of conditions involving modifications to the external resistance. As expected, reducing external resistance caused a drop in the initial ΔE value (from 3 V to 1.5 V) and a corresponding increase in the initial current density (from 7.6 A·m⁻² to 15.1 A·m⁻²). After 1 h, TOC removal rates were 31%, 57%, and 85% with external resistances of 38.5, 10, and 1.5 Ohms, respectively. The inherent resistance within a REDR consists of four main components: the area-specific resistances of the AEM and CEM, and the resistances of the compartments filled with LC and HC solutions. The REDR’s resistance is significantly influenced by the LC solution due to its essential role in ion transmission. Intensifying the LC solution concentration facilitates a reduction in the REDR’s resistance. As a direct consequence, there ensues a bolstering effect on Reactive Black B (RhB) eradication, hydrogen peroxide (H₂O₂) creation, and electric power production; these processes all exhibit growth concurrent with rising HC and LC concentrations [88].

However, at higher current levels, the mineralization efficiency of the REDR system decreases. This occurs because the production of oxidants at the electrode surface reaches saturation, leading to concentration polarization [73]. Additionally, surplus electron flow can trigger side reactions, notably oxygen evolution at the anode, which affects the system’s overall efficiency [72].

4.2. Degradation Circulation Mode

As previously noted, potent oxidizing agents are generated at both electrodes—the anode and cathode—in a REDR via electrochemical processes. These oxidants are instrumental in efficiently degrading organic contaminants present in wastewater, fulfilling a vital function in water purification systems. Two modes of organic wastewater degradation were implemented in REDR systems powered by salinity gradient energy: independent and synergetic circulation, as illustrated in Figure 4 [44].

As depicted in Figure 4A, the independent circulation mode utilizes twin wastewater reservoirs, each equipped with an exclusive hydraulic circuit. The anodic rinse solution and cathodic rinse solution function autonomously within their loops, degrading organic pollutants via specific anodic or cathodic reactions. Moreover, the reactants generated in these loops do not interact with each other. In contrast, the synergetic circulation mode incorporates a sole wastewater holding tank connected to a cohesive hydraulic system. Wastewater sequentially travels through both anodic and cathodic compartments before consolidating into a centralized tank, fostering potential synergistic interactions post-combination. As a result, the reactants from the anodic and cathodic reactions can either enhance or reduce the overall degradation efficiency.

During trials employing REDR for AO7 wastewater degradation (electro-oxidation via active chlorine at the anode and the EF process at the cathode), the average removal rates were 59% and 68%, respectively, when using independent circuits, outperforming the synergetic mode [91]. The lower TOC removal rates observed in the synergetic mode may be due to the interaction between H₂O₂ and HClO, which lowers the concentrations of both oxidants. Xu et al. [44] illustrated that, over equivalent treatment durations, the independent degradation mode exhibits superior AO7 dye decolorization efficacy relative to the synergetic mode. After 20 min of treatment, the independent mode showcased 99.93% decolorization in the anodic circuit and 96.52% in the cathodic circuit, contrasting sharply with the synergetic mode’s 82.85%. This disparity in performance can be attributed primarily to two key elements impacting the synergetic degradation circulation mode [44]:

(1)

Upon wastewater transit through both anodic and cathodic channels, residual oxidants may interact and mutually consume each other, thus reducing their effectiveness in contaminant removal. The relevant reaction equations are:

$${\mathrm{H}}_2{\mathrm{O}}_2+\mathrm{HClO}\to {\mathrm{H}}_2{\mathrm{O}+\mathrm{O}}_2+\mathrm{HCl}$$

(31)

$${\mathrm{ClO}}^{-}{+\mathrm{H}}_2{\mathrm{O}+2\mathrm{e}}^{-}\to {\mathrm{Cl}}^{-}+{2\mathrm{OH}}^{-}$$

(32)

(2)

In the context of EF processes, Fe²⁺, serving as essential catalysts, can undergo oxidation to Fe³⁺ at the anode. This transformation reduces the concentration of active Fe²⁺ ions, thereby diminishing their catalytic role and weakening the overall EF reaction efficiency.

According to the EF reaction equations at the cathode (Equations (16)–(19)), a portion of hydrogen ions (H⁺) is consumed during hydrogen peroxide generation, causing an increase in the pH of the cathodic rinse solution. To sustain the EF reaction in the cathodic loop, the continuous addition of diluted hydrochloric acid was necessary to maintain a stable pH of around 2.0 [44]. However, in the electro-oxidation process with active chlorine (Equations (25)–(27)), hydrogen ions (H⁺) are generated. In the synergistic degradation mode, hydrogen ions produced at the anode balance with those consumed at the cathode. Though decolorization might be less efficient compared to standalone modes, this method boasts a major advantage: eliminating the need for additional acids, thereby reducing treatment costs significantly. Despite slightly reduced color removal, synergistic degradation remains economically favorable by avoiding extra acid use.

Leng et al. [63,72] introduced an innovative combined process (AO&EF), designed to circumvent the interference between active chlorine species and hydrogen peroxide, which is common in traditional synergistic degradation loops. The novel methodology preserves a pivotal feature of synergistic degradation—the harmonization of hydrogen ions created at the anode and expended at the cathode. Consequently, this ensures a consistent pH level in the wastewater over the course of the entire treatment procedure, enhancing stability and efficiency. Additionally, the average energy consumption during degradation can be as low as 4 kWh·kg COD⁻¹ in experiments, which is significantly lower than that of conventional electrochemical degradation techniques. Consequently, the REDR system offers considerable cost savings in the treatment process, making it a viable option for adoption in the wastewater treatment industry.

Wang et al. [73] used a Ti/Ti₄O₇ anode and nickel foam cathode for the AO and EF processes in the REDR system. The oxidizing agents generated at both electrodes—hydroxyl radicals and adsorbed hydroxyl radicals—functioned without interfering with each other. Under a 0.2 A output current, the REDR showcased superior performance metrics: a COD reduction of 79.9%, a current efficiency of 79.6%, and an energy expenditure of 125.0 kWh·kg⁻¹ COD. These outcomes place the REDR ahead of conventional approaches to phenol-contaminated wastewater remediation, highlighting its enhanced effectiveness and efficiency.

4.3. Multi-Stage REDR-Series System

A stark salinity contrast between HC and LC compartments, post-saline solution thermal segregation, strains a single-stage REDR. The inefficient conversion of salinity energy to degradation power occurs due to poor mass transfer and limited surface area [93]. The membrane electric potential gradually decreases as working fluids move through the channels, primarily due to ion transport and mass transfer. In a REDR, ensuring that the electrode potential stays beneath the minimal theoretical potential at the exit point is critical. This precaution prevents the electrodialysis effect, a condition that leads to internal potential losses. To curtail energy wastage and enhance the capture of SGE, a multi-stage RED serial configuration is commonly adopted within REDHE [20]. Consequently, a multi-stage REDR-series system maximizes the extraction of SGE from the working fluids, enabling the effective treatment of organic wastewater. Xu et al. [20] conducted experiments simulating a series system to investigate azo dye degradation using a single REDR system, as illustrated in Figure 5 [20]. Within the multi-stage system, variations in salinity gradient influenced the output voltage of each REDR stage but showed negligible effects on the efficacy of color removal when treating azo dye-contaminated effluents.

Given the concurrent operation across multiple reactors, the multi-stage RED-series system showcases substantially enhanced degradation efficacy compared to a single REDR. In the multi-stage RED-series system, variations in output current and the number of reactors had a considerable influence over both contaminant breakdown proficiency and energetic transformation yields. Although increasing the output current enhances the decolorization efficiency for each reactor, this increase simultaneously causes a sharp reduction in the overall reactor capacity, thereby decreasing the cumulative decolorization rate for azo dye-contaminated wastewater. Fascinatingly, as the series-connected reactor array scales up, there arises a virtually direct correlation between its dimensionality and the aggregated AO7 azo dye decoloration throughput. Provided the system’s net output power is maintained above zero, the maximum decolorization rate of AO7 azo dye, reaching 1.91 mg·s⁻¹, is achieved with a current density of 0.10 A and a setup of nineteen reactors [20].

Leng et al. [53] formulated a computational framework applicable to a multi-stage REDR arrangement, aimed at emulating the remediation of effluents via the AO and EF processes. Numerical analyses reveal that the series system achieves the highest COD removal and peak power output at lower current levels due to its ability to incorporate more reaction chambers under low-current conditions. Additionally, the net power variation with the number of reactors follows a parabolic curve due to the increased pumping power required for more reactors. The electricity conversion efficiency of SGE increases with the number of reactors, reaching a maximum of 24% with 34 reactors. Wang et al. [60] investigated the energy-related effectiveness of a multi-stage REDR system for treating organic wastewater. Findings revealed that augmenting the tally of REDR modules integrated into the serial configuration facilitated greater harnessing of SGE, thus bolstering the ensemble electrical output. However, the number of REDRs (n) depended on both the output current and fluid velocity metrics, while the initial salinity gradient remained constant. An uptick in current draw or a downtrend in velocity led to a diminished REDR ensemble size. Moreover, experimental records unveiled that under parameters encompassing a 0.2 A current, a feed solution velocity of 0.5 cm·s⁻¹, and eight reactors, the system attained a maximum ΔCOD value of 450.98 mg·L⁻¹. The highest total energy conversion efficiency recorded was 11.41%, achieved with a current setting of 0.2 A, a feed solution velocity of 0.25 cm·s⁻¹, and three reactors.

5. Conclusions

This review highlights the recent breakthroughs in reverse electrodialysis driven by low-grade heat for wastewater treatment, with a primary focus on electrochemical oxidation degradation processes. Techniques, including cathodic reduction, electro-Fenton, anodic oxidation, electro-oxidation with active chlorine, and electrocoagulation, are seamlessly integrated at the REDR system’s cathode and anode. The amalgamation of electrochemical oxidation with reverse electrodialysis technology facilitates the efficient conversion of salinity gradient energy derived from low-grade heat into chemical energy. This conversion is instrumental in degrading pollutants, effectively addressing the treatment of toxic Cr(VI), dyeing wastewater, ammonia-rich wastewater, and natural water bodies, thereby offering a sustainable solution to environmental challenges.

In the REDR wastewater degradation process, several strategies can boost the output current density, including the augmentation of membrane pairs, enhancement of HC and LC flow rates or salinity gradients, and reduction in REDR resistance. These enhancements accelerate electrode reactions, generating more oxidants for pollutant degradation and improving wastewater degradation efficiency. Oxidative degradation at the anode and cathode can occur independently in the independent degradation circulation mode or synergistically in the synergetic mode, where both electrodes work in concert for wastewater treatment. In combined degradation, anodic and cathodic oxidative processes can either inhibit or enhance each other, contingent on whether intermediate products from oxidation impact the opposing electrode’s degradation process. Compared to a single REDR system, multi-stage REDR series allow for more extensive SGE extraction from working solutions, leading to faster wastewater degradation rates.

While the incorporation of SGE in REDR for wastewater treatment offers clear advantages, such as environmental friendliness and sustainable performance, numerous challenges remain in transitioning RED technologies from experimental prototypes to scalable industrial applications. Primarily, the current high cost of RED membranes and electrode materials significantly increases the operational expenses of RED-driven wastewater treatment processes [12]. Addressing this cost barrier is essential to facilitate broader adoption and commercial viability. Additionally, the efficiency of salinity gradient energy transformation in RED systems lingers at modest levels. A significant portion of the energy is lost due to high internal resistances [28,41], resulting in inefficient use of available salt gradients—an issue that urgently requires attention and innovative engineering solutions to improve energy recovery. Most critically, the inherent nature of RED technology limits its applicability mainly to regions with abundant salinity gradients, such as marine–freshwater interfaces [94]. This geographical limitation severely restricts the deployment of RED techniques in inland areas where salt gradients are scarce.

Addressing the aforementioned challenges, future research endeavors will concentrate on pivotal domains:

• Developing new materials and membrane technologies to enhance SGE conversion efficiency and durability.
• Investigating design and optimization strategies for integrated systems to maximize energy conversion efficiency while minimizing operational costs.
• Evaluating the economic feasibility of SGE in various applications, including comparisons with conventional energy and desalination technologies.

Through these research efforts, the REDHE system driven by low-grade thermal energy can achieve higher degradation efficiency and energy conversion performance. The integration of low-grade waste heat significantly reduces the energy consumption of the REDHE system, making it a more cost-effective and sustainable solution for industrial wastewater treatment. This breakthrough not only addresses environmental challenges but also contributes to the sustainable development of energy and water resource management.