Fabricating EM preparation involves preparing membrane active layers with

elevated electrical conductivity and porosity. Sun et al. [1] revised the literature from the previous decade on conductive materials employed for EM fabrication, comprising carbonaceous materials, metals, metal oxides, and polymers.

The large part (about 60%) of the EMs investigated are fabricated employing carbonaceous materials (Figure 2(A)) [1]. In such materials, the carbon atoms are mainly sp² hybridized with delocalized π-π electrons, spreading elevated conductivity. Carbonaceous materials can work as either the anode or the cathode. Most carbonaceous anodes are viewed as “active” anodes with an overpotential for oxygen evolution reaction generally lower than 0.4 V [1]. The “active” anode surface (M) interacts highly with electrogenerated hydroxyl radicals (^•OH) to produce a higher anodic oxide, as shown in Equation (1), which is only apt of partially oxidizing organic substances. Thus, a voltage below the oxygen evolution potential (OEP) of carbonaceous electrodes is usually applied for usages implying physicochemical-founded pathways (e.g., electrostatic repulsion, electrophoresis, and direct oxidation/reduction) [23].

M ( • OH ) → MO + H + + e − (1)

Carbon nanotubes (CNTs) remain the most frequent conductive materials for EM preparation thanks to their applicability for facile generation of an interwoven

and compact structure with high electrical conductivity and porosity [24]. Such structure authorizes CNTs to bond with each other via van der Waals interactions, that way keeping their stability throughout EM process [1]. Graphene-based materials, comprising graphene oxide and reduced graphene oxide, as well show useful electrical features. The conductive layer may be synthesized by deposition, conducting to an elevated specific surface area and controllable pore size distribution [25]. Functionalization with oxygen-containing groups authorizes CNT- and graphene-based materials to bind with other metal catalysts (e.g., Fe and Pd) [26]. Boron-doped diamond (BDD), one of the most encouraging flat plate “nonactive” anodes for ^•OH formation (i.e., anodes with elevated OEPs) [27], stays scarcely mentioned for EM preparation, since the techniques for preparing the BDD layer with elevated specific surface area and porosity remain complicated and expensive [28]. Further, mesh electrodes (like carbon fiber cloth and carbon paper) may be utilized as conductive substrates for the modification of membrane structures [1]. The usage of high specific surface area coal and graphite has as well been noted [29].

Metals and metal oxides are an additional class of conductive materials that remain encouraging for EM fabrication. Among such materials, several studies have been dedicated to the employment of Magnéli phase Ti₄O₇ [30] [31] [32], the most conductive phase of the substoichiometric titanium oxides (Ti_nO_2n−1, 4 ≤ n ≤ 10) [33]. When used as a “nonactive” anode, the OEPs of Ti₄O₇ (2.2 - 2.7 V vs SHE) are comparable to those of BDD electrodes (2.3 V vs SHE) [34], showing its superior efficiency for ^•OH formation through water oxidization. In addition, Ti₄O₇, having a porous, stable monolithic structure, may be prepared via anodic polarization [1]. Among metal materials, Ti mesh remains one of the most encouraging metal substrates employed for functionalization with electrocatalysts (like IrO₂, RuO₂, and doped SnO₂) [35]. Further, steel mesh substrates may act as cathodes for electrochemical reduction or for catalyzing electro-Fenton reactions [36].

Conductive polymers have a conjugated backbone that generates a series of alternating single and double carbon bonds [37] [38]. The overlap of p-orbitals in the π-bonds authorizes the electrons to delocalize and move between atoms. Polypyrrole and polyaniline have been employed largely for EM preparation thanks to their stability under harsh conditions (like high pressure and chemical exposure) [1]. Nonetheless, several polymer-based EMs suffer from low conductivity and water flux [38]. Integrating conductive polymers with carbonaceous materials (e.g., CNTs and graphene) could ameliorate the electrical efficiency of EMs.

EMs are usually fabricated utilizing free-standing membrane casting, membrane modification, or composite membrane assembly [1]. Free-standing Ti₄O₇ and carbonaceous EMs, having both separation and electrical features inside a porous monolithic structure, may be fabricated through the membrane casting method [39], as depicted in Figure 2(B₁). Frequently, a mixture of the conductive materials and binders (e.g., paraffin oil and polymers) [1] is first synthesized. The free-standing EMs are prepared through mechanical pressing, gel casting, or wet-spinning methods to shape the mixture, followed by thermal treatment to remove the binder.

The most usual technique for EM fabrication consists of modifying a porous substrate with conductive materials (Figure 2(B₂)). Conductive substrates (e.g., carbon filter cloth, Ti mesh, stainless steel, and PANI) remain largely utilized. Functionalization of electrocatalytic materials (e.g., CNTs, graphene, and doped SnO₂ on the substrates) ameliorates their electrochemical efficiency and/or controls the membrane pore size. Cross-linked polymers with conductive materials like CNTs could authorize accurate control of the EM pore sizes. Further, using nonconductive substrates, comprising ceramic (Al₂O₃ and TiO₂) and polymeric (polyvinylidene difluoride and polytetrafluoroethylene) membranes is as well frequent [1].

A straightforward procedure to fabricate EMs consists of assembling mesh electrodes on commercial membranes (Figure 2(B₃)). In such situation, flat plate electrodes with big pore sizes can be employed as the active layer. A mesh electrode could be pasted onto the membrane substrate or assembled into an electrofiltration module with a membrane [40]. Sun et al. [1] noticed that the present performance of these EMs carrying conductive layers with large pore sizes stays much lower than that of EMs fabricated by the other two procedures, even if their specific surface areas can be comparable, as discussed later.

Various types of flat-sheet, hollow-fiber, and tubular EMs could be employed for water treatment [41]. EM modules and filtration modes are shown in Figure 2(C₁), Figure 2(C₂). In usual usage, the EM frequently works as one of the electrodes, whilst the counter electrode is placed either on the feed or on the permeate side. The conductive layer of the EM usually faces the feed side [1].

Cross-flow filtration remains the most frequent mode for practical water treatment usages. In such filtration mode, the feed flow passes along the membrane surface and the EF is perpendicular to feed flow (Figure 2(C₁), Figure 2(C₂)). Dead-end filtration, in which feed flow is perpendicular to the membrane, suffers from rapid formation of a cake layer, which hinders EM efficiency [1].

Benefiting from the flow-through configuration, EMs may reach enough elimination of pollutants via convection-enhanced mass transport [1]. Traditional nonmembrane electrified techniques, run in batch or flow-by modes utilizing flat plate electrodes, conduct to the generation of thick diffusion boundary layers (~100 μm). Consequently, such techniques suffer from mass transport restrictions, since the overall reaction rate is governed by the diffusion rate of the pollutants to the electrode surface (Figure 2(D₁)) [42]. By contrast, in flow-through mode, the usage of porous electrodes like EMs diminishes the thickness of the diffusion boundary layer to a length scale that is comparable to the membrane pore radius (Figure 2(D₂)), therefore considerably ameliorating the mass transport rate. Usually, the smaller the EM pore size and the higher the water flux, the lower is the mass transport restriction [1].

When the membrane pore size is on the scale of nanometers (i.e., UF scale), the pollutant elimination may be more elevated via nanoconfinement [43]. The main pathways induced by nanoconfinement involve the enrichment of reactant species, like ^•OH, in nanoconfined spaces, as well as the decrease of activation energy barrier or change of reaction pathways and kinetics [44]. Carbonaceous materials (like CNTs), which are frequently employed to produce porous structures with nanoconfinement, are as well efficacious for the adsorption of target pollutants within the membrane pores, augmenting the residence time and reduction performance [1]. The adsorption rate and capacity could be enhanced considerably via electrochemical assistance [45]. Nanoconfinement as well enriches the local level of pollutant molecules, encouraging adsorption and enhancing the following reactions [46].

The pathways examined in this Section could authorize EMs with outstanding features to attain a single-pass elimination of pollutants in seconds [1]. In particular, the elimination rates are several orders of degree greater than those attained by traditional batch and flow-by modes employing flat plate electrodes [47].

EMs eliminate viruses via combining physical retention with EF engendered electroadsorption and electrooxidation (EO) [58] [59] [60] (Figure 3(A)). Using EMs as the anode electrically adsorbs negatively charged viruses onto the membrane surface throughout filtration [61] [62] [63], while concurrently demobilizing viruses via producing oxidative stress by direct and indirect EO [64] [65] [66]. Also, reactive oxidants like ROSs and RCSs are diffuse in the permeate, stopping viral diffusion and securing water safety [1]. Oxidant production rates are crucial to virus demobilization in EM filtration, which is considerably affected by membrane materials, applied voltage, and feedwater chemistry [67] [68].

CNTs are recommended for preparing anodic EMs over a set of conductive materials like titanium suboxide and silver nanowires [69]. Without using an external voltage, a CNT-EM has attained 4 log virus elimination (with an initial virus concentration of 10⁷ plaque-forming units (PFU)/mL) via aromatic π-band

induced adsorption. When an external potential (2 and 3 V) was used with the EMs, no culturable viruses (>6 log elimination) could be observed in the permeate [70]. Viruses are demobilized in electrofiltration with a residence time (in CNT apparatus) of just 30 s [70], probably due to 1) the destruction of viral morphology under EF [70] and 2) the produced oxidative stress on viral envelope, capsid protein, and nucleic acids provoked by ROSs/RCSs [71]. Juxtaposed with traditional electrochemical techniques for point-of-use virus demobilization [72] [73] [74], the energy consumption of the CNT-EM is lower: just 8.3 Wh/m³ of permeate (with applied voltage of 2 V) [1].

Practical solution chemistries of feed waters influence not only electrochemical reactions (like oxidant formation) but also viral migration and aggregation behavior, as a result of that affecting the viral demobilization efficacy of EMs [1]. Three well-examined solution components influencing viral demobilization are natural organic matter (NOM) [75] [76] [77], chloride ions [78], and divalent cations (Figure 3(B)) [79] [80]. Co-occurring NOM in feedwater probably influences disinfection performance of EMs via 1) competing with viruses for oxidant consumption, 2) stopping the agglomeration of viral particles through electrostatic stabilization, and 3) averting virus-membrane contact via blocking EM reactive sites or binding with viruses to form a shielding layer [70]. Chloride ions considerably enhance virus demobilization effectiveness of a CNT filter at an applied voltage over 2.3 V, since such potential is apt to initiating the one-electron oxidation of Cl⁻ to RCS [70]. Even if the RCS production is beneficial for viral elimination in EM process, the hazard of disinfection by-products (DBPs) formation remains an issue [81] [82] [83]. As a result, DBPs level has to be constantly observed when high-level chloride ions coexist with viruses in the feedwater [84]. Divalent cations (e.g., Ca²⁺) conduct to virus agglomeration via charge neutralization [85] or by bridging between viral particles and NOM (or CNTs), and then improve virus elimination via EMs by physicochemical filtration [80].

In addition to EMs, additional types of membranes (like nanocomposite membranes and flow-through electroporation) have been tried for disinfecting viruses [1]. Nanocomposite membranes functionalized with metal nanoparticles (NPs) (e.g., AgNPs, Cu₂O, TiO₂NPs, Ag-/Cu-TiO₂ hybrid NPs) manifest considerable viral elimination (3 - 5 log) via synergy among physical retention by the membrane and the biocidal property of metal NPs (Figure 3(C)) [86]. Since the effectiveness of nanocomposite membranes mostly depends on contact among viruses and anchored NPs, low permeate flow rate (permeate flow rates of 10¹ - 10² L/(m²h) are needed for viral disinfection [1].

Electroporation uses sharp tip structures (like metal nanowires) on a membrane electrode to locally improve the EF [1]; this creates a transmembrane potential that conducts to pore generation in the lipid bilayer of enveloped viruses and following viral inactivation [87] [88]. Even if electroporation offers superior bacteria disinfection (>6 log bacterial removal under a flow rate of 10³ - 10⁴ L/(m²h) (Figure 3(C)) [89], its virus elimination (1 - 2 log) is notably lower under identical circumstances. Further, disinfection through electroporation of nonenveloped viruses has not yet been studied. Among all types of membrane, EM filtration attains similar efficiencies for both virus and bacteria disinfection under a flow rate of near 100 L/(m²h) with moderate applied potential (2 - 3 V) [1]. The comparatively long retention times required are a great obstacle, since these cannot be attained in practical membrane filtration without recirculation.

Changing membrane-electrode geometry could increase viral disinfection performance. Narrowing membrane channels creates nano/micro-scale spatial confinement in EM filtration, encouraging electrochemical reactions via augmenting the frequency of collisions among viral particles and electrically produced oxidants [1]. Exploratory and computational investigations have established that patterned nanostructures ameliorate mass transport in the electrode-solution interfacial region, probably via the next pathways (Figure 3(D)): 1) nanostructures locally boost the EF that grows oxidant production on electrodes [90]; 2) patterned nano-geometry leads to nano/microscale turbulence, augmenting convection and local mixing of reactants [1]; 3) the geometric morphology conducts to spatially confined redox reactions at electrode/solution interfaces that produce temperature gradients along the nanostructures, which then speed up local convection and improve reactant mobility. Nonetheless, the feasible usage of such approach faces dares about the stability of patterned nanostructures under complex water matrices and the preparation of the sophisticated nanostructures inside porous EMs. Novel materials and techniques to manufacture nanostructure engineered EMs, which could furnish a large surface area, elevated conductivity, and increased corrosion resistance, deserve more study. Moreover, detailed techno-economic studies must be performed to evaluate the economic feasibility of such techniques contrasted to traditional desalination processes.

Even if numerous investigations have been dedicated to EMs for virus demobilization, EMs engineered just for such objective are far to attain economic feasibility as techniques implying chemical disinfection or membranes with smaller pore sizes (such as “tight” UF, NF, and low-pressure RO) could already attain virus disinfection/removal. In MF/UF techniques, virus demobilization can be simultaneously reached when engineering EMs for water treatment through electrochemical oxidation. Therefore, a large number of the mentioned dares for electrochemical oxidation are as well applicable to dares particular to virus demobilization. As an illustration, examining parameters of EMs for virus demobilization should extend beyond a single model virus and ideal water matrices because of the influences of species variations and the occurrence of NOM integrated with low virus levels. Also, nanostructure-enabled EM techniques for virus demobilization mostly lack practicality, due to the sophisticated processes for material preparation and the unpredictable operational stability under an EF and harsh water flow circumstances [1].

Even if the working concept of ED through in situ formation of oxidants does not change, the specific oxidant formed may influence the needed dose across treatment contexts and introduce chemical reaction pathways that compete with pathogen demobilization [108] [109] [110]. Besides, the water composition across treatment contexts, and then the oxidant demand for disinfection, may change greatly (Figure 4). For ED, the three frequent treatment contexts are water treatment, centralized wastewater treatment, and distributed wastewater treatment [91]. In the water treatment context, ED may take place following coagulation/flocculation, sedimentation, and filtration stages where relevant, analogous to conventional chemical disinfection methods [91]. Whereas the scale of centralized water treatment systems could change, the composition and quality of the water to be treated depend on source and season. Within the centralized wastewater treatment context, ED would also proceed analogously to conventional chemical disinfection methods following primary treatment and secondary biological treatment at a centralized treatment plant.

Novel investigations focused on augmenting sanitation access throughout the world and ameliorating present sanitation system resiliency has conducted to

refreshed attention in ED setups for treating and disinfecting water [111] [112] [113]. Nonetheless, many investigations have tested ED device efficacy, reporting metrics have varied largely [114] [115]. When assessing ED devices both internally (like across oxidant types, working parameters, and setup arrangements) and externally (such as conventional chemical disinfection, filtration techniques [116] [117] [118], and thermal demobilization [119] [120] ), it is crucial to evaluate both the disinfection performance and the energy consumed throughout process [91].

As a consequence, next researches with electrochemical treatment must directly detail both the dose produced and the energy consumption throughout production to let for juxtaposition across working parameters and procedures [91] [121]. If dose cannot be directly estimated, as in the instance of several ^•OH founded devices, then the concentration of oxidant must be evaluated and detailed utilizing regularly adopted practices [122] [123]. Also, the charge efficacy of oxidant formation must be detailed where possible to contrast running features across the range of treatment contexts [124].

Hydroxyl radicals seem to be an encouraging oxidant for application in electrochemical treatment thanks to their high oxidant strength and possibility for demobilizing highly recalcitrant pathogens like Cryptosporidium spp. [91]. Nevertheless, their efficiency has not been as easily assessed comparatively to more usual oxidants (e.g., Cl₂) [125] [126] [127]. Thus, more investigations remain required to both describe the efficacy and navigate trade-offs among energy consumption and oxidant strength in ^•OH founded setups, especially at adequate doses for killing viruses and Giardia. By more comprehending disinfection effectiveness and energetics in ^•OH founded devices, the comparative strengths and demerits of various oxidants could be defined for applicability across treatment contexts. Within resource limited settings, merging renewable energy sources (like photovoltaic cells) constitutes a significant route for sustainable ED [128] [129] [130].

Finally, substitutional arrangements and functioning procedures stay to be explored particularly targeting energy per dose lowering [91]. Both empirical investigations and simulated efficiency indicate that using lower cell voltages through prolonged residence periods conducts to appreciably lowered energy per dose (Figure 5). In both centralized and distributed treatment contexts, this will probably need engineering vessels that maximize the residence period over disinfection. More investigations stay requested to determine and navigate expected trade-offs among effectiveness and economic viability of batch, semi-batch, and continuous operation when used at high residence periods (especially for larger, centralized systems that usually work at very high throughput and can need continuous treatment setups). Even if such modifications could constitute an unsuitable handicap for present centralized treatment systems, prolonged contact periods are habitual in the batch or semi-batch treatment systems largely employed for distributed or semi-centralized treatment in resource-limited settings [131] [132]. Considerable dares endure for the diffuse deployment of robust ED setups. Nonetheless, when employed inside treatment contexts that boost intrinsic advantages, ED offers real promise for increasing worldwide access to sanitation and safe drinking water (Table 1).