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Review

Novel Materials for Heavy Metal Removal in Capacitive Deionization

by
Youze Xu
1,
Zhenyu Zhong
1,
Xianhui Zeng
2,
Yuanyuan Zhao
1,
Wenting Deng
2 and
Yuehui Chen
2,*
1
Key Laboratory of Water Pollution Control Technology, Hunan Province, Hunan Research Academy of Environmental Sciences, Changsha 410004, China
2
School of Environmental and Resources, Xiangtan University, Xiangtan 411105, China
*
Author to whom correspondence should be addressed.
Appl. Sci. 2023, 13(9), 5635; https://doi.org/10.3390/app13095635
Submission received: 7 March 2023 / Revised: 17 April 2023 / Accepted: 20 April 2023 / Published: 3 May 2023
(This article belongs to the Special Issue New Frontiers in Separation of Heavy Metals: Removal from Industrial Wastewaters and Contaminated Soil)
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Abstract

:
Heavy metals are considered a class of contaminant that can accumulate in the food chain and thus must be removed from contaminated media. Heavy metals can be removed by electrocoagulation, electroflotation, electrodialysis, capacitive deionization, and so on. Among the methods to remove heavy metals, capacitive deionization is one of the most attractive methods that can remove heavy metal ions without using a large volume of chemicals and producing a high number of heavy metals containing solid wastes. In this study, after a brief introduction to the mechanism of capacitive deionization, we focus on materials that have been developed as electrodes for heavy metal removal in capacitive deionization and summarize the latest advancements. Finally, with particular emphasis on material design, we provide some further insights in this area.

1. Introduction

Industrialization brings about the rapid transformation of manufacturing and leads to higher productivity, raising living standards. However, this process also brings threats to human beings such as the pollution of air, water, and soil, which can result in significant deterioration of quality of life and life expectancy [1]. Heavy metals (HMs) are regarded as one of the most important categories of pollutants and have caused wide concern due to their high toxicity and bioaccumulative properties. The bioaccumulative process leads to the pollution of air, water, and soil and eventually leads to adverse effects on humans. HMs have no rigorous definition on a scientific and chemical basis [2]. Generally, HMs have a density of over 5 g cm−3 (5 times >H2O), meaning a metal with a density below this lacks heavy metal chemistry. Thus, HM properties are partially related to density as well as their chemical properties; they contain transition and post-transition elements along with metalloids, namely arsenic and selenium. Moreover, HMs are regarded as “toxic elements” priority contaminants listed by the United States Environmental Protection Agency [3]. A large number of HMs, with the exceptions of Cd, Hg, and Pb, are also essential micronutrients for living organisms; they only become toxic when the concentration is over the limit [4,5]. The dosage that determines either HM toxicity or micronutrient properties is the critical criterion with which to identify the toxicity of HMs [6].
Sources of HMs can be natural such as landslides, weathering, and volcanic eruptions, but they mainly originate from anthropogenic processes such as mining, agriculture, electroplating fossil fuel combustion, and waste disposal. For instance, agricultural use of HMs in the form of pesticides and fertilizer as well as corrosion and industrialization releases HMs into the environment [7,8]. Untreated wastewater containing heavy metals discharged into the natural environment can pollute water bodies and soil, endangering food safety and human health through the food chain [8]. An increase in concern for public health has led to the enaction of varied legislation to control the emission of HMs [9]. This increase in public health concern has led to the enaction of varied HM emissions control legislation. This leads to an eruption of interest in developing new technologies to control emissions. Widely applied technologies for HM removal are based on solid-phase extraction using precipitation, coagulation, and flocculation [10]. However, a major concern with this technology is its output of heavy metal solids, which become solid waste. Conversely, liquid-phase extraction based on new technologies such as adsorption, ion exchange, and membrane separation has gained much interest because these processes concentrate metal ions directly devoid of a huge amount of solid heavy metal-containing waste. However, new advanced liquid phase extraction methods are highly dependent on emerging new materials such as adsorbents and ion exchangers, membranes, and electrochemical adsorption materials. [11,12]. In the past several decades, great efforts have been made toward developing new materials for heavy metal removal. With the exception of adsorption and membrane filtration processes, electrochemical treatments for HM removal are gaining increasing interest. These methods use electrostatic fields to dry metal ions and enhance precipitation, adsorption, and separation. And this paper reviews the electrochemical treatment technology of heavy metal wastewater, which can provide a reference for the development of efficient and energy-saving heavy metal wastewater treatment technology. Figure 1 summarizes electrochemical methods for HM removal. These electrochemical methods can be classified into electrocoagulation, electroflotation, electrodialysis, and electrodeposition depending on their mechanism.

2. A Brief Review of Electrochemical Methods

Electrocoagulation (EC) is currently an efficient technology in wastewater treatment [13]. EC is a process involving the in situ generation of coagulants such as aluminum/iron ions and corresponding hydroxides from the electrical dissolution of “sacrificed anodes” under direct electrical currents [14]. Generated iron or aluminum flocs serve as coagulants and subsequently precipitate HMs, and an aluminum plate, steel/iron plate/mesh, and Ti plate/mesh are applied as anode and cathode materials to remove Cr(VI), Cu(II), Ni(II), Cd(II), Pb(II), Zn(II), Ba(II), and As(III) [15,16,17,18,19,20,21,22,23]. The EC process can dispose of multiple contaminants in one process; has low sludge production and maintenance costs; can easily handle water quality variations; and can achieve over 95% removal efficiency for HMs. [15,18,22]. The disadvantages of EC are as follows: (1) The dissolution of “sacrificial electrodes” into wastewater occurs due to oxidation, which requires frequent disposal. (2) The passivation of electrodes over time has limited its implementation. (3) The use of electricity may be expensive in many places. (4) High conductivity of the wastewater suspension is required [24,25,26].
Electroflotation (EF) is suitable for separating complex heavy metal wastewater with heterotic contaminants such as sludge. In this process, water is split into oxygen and hydrogen electrodes. Gas bubbles are generated, and the suspended contaminants are raised to the water surface. In this complicated process, bubble size and gas generation rate are two crucial factors in metal efficiency, deeply dependent on the current density used [27]. EF is gaining interest because its system produces a fine gas bubble without turbulence, the bubbles exhibit good coverage of the whole area, and the system is suitable for full automatization. The EF process has been applied in the food industry; oily wastewater; chemical wastewaters; effluents from the sugar and leather industries; and farms as well as in the dairy industry, and thus has been successfully commercialized [28]. Gas bubbles are the workhorse of EF, and bubble size is the key parameter for determining efficiency. Recently, a coupled method combining EF and EC to remove and separate heavy ions has been scaled up in industry [29]. In such a process, metal ions are coagulated and floated over the water surface [30]. The drawbacks of EF are the insufficiently high productivity of its units; the release of H2 bubbles; the cost of electrodes and maintenance; and the formation of sludge by volume [31].
The EF system relies on hydrogen and oxygen/chloride produced at the cathode and anode, respectively, which are referred to as the hydrogen evolution reaction and oxygen/chloride evolution reaction, respectively. These processes are particularly energy-consuming due to the water-splitting process, requiring a potential over 1.5 V even when using the best-developed catalysts. Presently, the most widely applied anode materials in EF systems are stainless steel, nickel plates, titanium, and carbon cloths due to their cheap price and relatively good HER ability [32,33]. Anode materials should be chemically stable since the electrodes are working at high potentials (1–3 V). Iron, aluminum, and stainless steel are cheap but suffer from corrosion [34]. Graphite and other carbon-based materials are relatively stable but exhibit high overpotential for OER [35]. Dimensionally stable anodes (DSA) are the most attractive anodes for EF [36,37]. The electrodes are developed using conductive precious RuO2, IrO2, and their composite metal oxides as the electrocatalysts, normally using TiO2, Nb2O5, Ta2O5, and ZrO2 as the stabilizing agents supported on a Ti-, Ta-, W-, etc. -based metal substrate [37,38]. The major merits of these materials are their high stability and long lifetimes even in low-pH solution; thus, these anodes provide an attractive alternation for EF.
Electrodialysis (ED) is used to transport metal ions between two solutions along ion-exchange membranes under the influence of an applied electric potential difference. ED is an established technology in treating industrial wastewater; wastewater from the drug and food industries; table salt production; chemical processes; and HMs removal [22,39]. It is conducted in a configuration called an ED cell [40]. The ED process can purify water without phase change, reaction, or chemicals [41]. These merits provide environmental benefits without the use of fossil fuels or chemical detergents. However, ED technology has limitations, including scaling, membrane fouling, energy consumption, and selectivity [42,43]. ED has been investigated for the removal of Cr(VI) [44,45], Cu(II) [46], Zn(II) [47], Ni(II) [48], and Ag(I) [49]. ED systems necessitate excellent ion exchange membranes with lower electrical resistance, mechanical durability, improved permeability, and higher selectivity for specific ions. Production costs have also come under consideration [32], which is discussed in this paper’s section on membranes.
Electrodeposition (EP) is a practical method that relies on the redox reduction of anodes alone. Heavy metal ions are reduced and deposited on an electrode in solid form. This process allows the simultaneous removal and recovery of HMs from wastewater [50,51]. High removal and recovery efficiency have been reported for the disposal of As(V)-, Cr(VI)-, and Cu(II)-containing wastewaters [51,52]. However, the reaction degree and side reactions such as water splitting were hard to control due to high energy dissipation [25]. Table 1 summarizes the application of EC, EF, and EP for heavy metal removal.
Capacitive deionization (CDI) is among emerging separation technologies for removing charged ions from wastewater using electrochemical principles on electrode surfaces. Figure 2 shows the mechanism of CDI for HM removal. Ions are removed from the aqueous solution and are stored on the internal surface areas of the porous electrodes; this results in an effluent product stream with reduced ion concentration. Moreover, ions can be reversed into solution by reducing or even reversing the applied voltage, resulting in a product stream concentrated in ions. In principle, CDI is based on the electrosorption of cations and anions on the electrode surface captured by electrical double layers [77]. Generally, the electrical properties of electrodes govern the adsorption capacity, chiefly specific capacitance, and their surface areas are believed to be closely related to the removal efficiency of CDI cells [78]. The CDI process may specifically remove heavy metal ions using an electric field without adding any chemicals, making the process an energy-efficient technology compared with distillation and reverse osmosis. In recent decades, CDI has grown substantially, considering electrode materials, cell configurations, and process analysis, while parasitic reactions and a relatively low desalination capacity hinder the practical application and further development of CDI [79]. Its side reactions mainly include anode oxidation reactions (such as carbon electrode oxidation and chloride oxidation), cathode reduction reactions (such as oxygen reduction and carbon electrode reduction), and co-ion expulsion (-i.e., ions with similar charge to the electrodes are expelled from the electrodes during the charging process in CDI). These reactions often occur during the operational process, increasing energy loss and leading to water quality fluctuations. The low desalination capacity is mainly due to the relatively low physical charge storage capacitance of the electrode materials and the adverse effects of side reactions. CDI technology is energy efficient, has a low life-cycle cost, is capable of removing a wide range of ionic contaminants, has a high recovery rate, and is cost-competitive compared with membrane filtration processes such as reverse osmosis (RO), which can be especially beneficial for saline groundwater treatment for rural or remote communities [78,79,80].

3. Materials for Heavy Metal Removing Using CDI

Great efforts to solve the challenges of CDI have been made. These include CDI configuration optimization and developments in operation style as well as the search for highly efficient and cheap electrode materials with high adsorption capacity and stability [81]. The electrode material is a crucial component of CDI cells that determines the desalination performance of CDI [78]. Usually, electrode materials should be conducive to a high surface area and have a good affinity with heavy metal ions. A large number of electrode materials have been developed for CDI. Figure 3 summarizes the materials applied for CDI, the majority of which include carbon materials, metal oxides, polymers, and their composites.

3.1. Pure Carbon Materials

Generally, adsorption capacity is governed by the electrical properties of electrodes, especially specific capacitance, and surface areas is believed to be relevant to the removal efficiency of CDI cells [82]. Carbon has abundant allotropes and has been extensively studied in the past several decades for a variety of applications. Owing to its high surface area, excellent conductivity, and tunable structure, carbon materials are the most popular materials applied for CDI electrodes. Large surface areas and mesoporous carbon materials exhibit good characteristics and excellent performance regarding electrochemical activity, the adsorption/sorption of toxic gases, and the removal of HMs [83].
Carbon materials can reversibly adsorb and desorb significant amounts of HMs using double-layer mechanisms without affecting their mechanical and electrical properties. Specific surface areas and pore sizes were found to closely correlate with ion electroadsorption at the electrical double layer [84]. High-surface-area porous electrodes must be optimized for both high surface area (capacity) and pore size (kinetics), and electrosorption performance was found to be primarily determined by hydrated size, ionic charge, and initial molar concentration in solutions. Nanoporous carbons are used as a key component in the electrosorption process due to their good electrical conductivity, high surface area, and remarkable sorption capacity. Carbon aerogels, activated carbons, carbon nanofibers, carbon nanotubes, graphene, mesoporous carbon, and carbon-based composites are recognized as promising electrodes for CDI [78]. Since the first report on the electroadsorption of Cr(VI) using carbon aerogel electrodes in 1997, increasing attention has been paid to developing new carbon-based materials for heavy metal electroadsorption [85].
Activated carbon is a notable material used for electrosorption because of its strong mechanical stability; large pore volume; high surface area; electrochemical stability; and low manufacturing cost. Many investigations have revealed that activated carbon is effective in removing ions such as Na, Mg, Ca, Cl, etc., from water. For example, in one study, the removal efficiencies were 32%, 43%, and 52% for Cd2+, Pb2+, and Cr3+, respectively, initially formed heavy metal-containing water with a concentration of 0.5 mM [86]. Electrosorption rates and ionic hydrated radii were not linear for heavy metal but in the order Cu2+ > Pb2+ >Cd2+, attributed to the complex mechanisms occurring in electrosorption, such as electrodeposition, and OH- complex formation during charging [86]. Cu2+ and Pb2+ showed notably low desorption efficiency during electrode regeneration, implying high fouling potential during CDI treatment because of the facial deposition of Cu (+0.34 V) and Pb (−0.13 V). Chen et al. investigated the electrosorption and reduction behavior of Cu2+ and Zn2+ in a wide voltage window on activated carbon fiber electrodes [87]. It was shown that removal of Cu2+ worked better than that of Zn2+ at 0.4–0.6 V, while better removal efficiency for Zn2+ occurred in the range 0.8–1.2 V in wastewater containing a single metal ion. During the membrane CDI process, Cu2+ could be preferentially electrodeposited while Zn2+ reduction was prevented at the voltage of 0.8 V [87]. As(III) and As(VI) removal action is known, and the removal capacity of activated carbon electrodes strongly depends on applied voltage and initial arsenic concentration. The sorption capacity for treating As(V) is superior to that for As(III) due to the negative charges of prevalent As(V) species. The removal of As(V) occurs mainly via electrosorption, whereas As(III) removal using the CDI system involves the oxidation of As(III) to As(V) before it is electrostatically adsorbed on the anode surface [88]. The fundamental behavior of Cu2+ removal in the CDI process over activated carbon electrodes has been investigated [89]. It was shown that electrical double-layer charging assisted Cu2+ removal, and electrodeposition can be curbed at a low applied voltage. The equilibrium electrosorption capacity of Cu2+ was 24.57 mg g−1 by the electrical double layer at a charge potentially less than 0.8 V, while electrodeposition of Cu was seen when the potential was over 0.8 V.
The adsorption capacity of electrodes also depends on co-existing organic and inorganic pollutants. Cr(VI) electroadsorption was raised from 155.7 to 190.8 mg g−1 when KCl concentration inclined in an electrolyte in the range of 100–500 mg L−1, but notably declined to 90.2 mg g−1 when the KCl concentration reached 1000 mg L−1. High Cr6+ removal efficiency occurred in the absorption of Cr(VI) onto electrodes, the reduction of Cr(VI) into Cr(III), and the formation of precipitation onto the electrode surface [90]. Co-existing Sr and Cs ions in aqueous solution largely inhibited the removal efficiency of Co ions from 36.54% to 8.37% [91]. Wang et al. used a carbon aerogel as the electrode for the CDI removal of Cu2+ [92]. The 3D network-like carbon structure was composed of small particles with mesopores and macropores having a pore volume of 3.41 cm3 g−1 nm in the carbon materials and had a great impact on the adsorption of diverse ions and the anion adsorption effect in the following order: SO42− > NO3− > Cl−. The highest specific capacity of 156 F g−1 in 6 M KOH solution, reaching a Cu2+ adsorption rate of 29.7 mg g−1, was achieved in an optimized condition in which the Cu2+ concentration was 100 mg L−1 and the applied voltage was 1.2 V.
Graphene is a two-dimensional material with a high specific surface area and intrinsic electrical conductivity along with mechanical robustness. It is considered an excellent candidate for CDI electrodes. Graphene-based materials have been widely studied for water desalination [93,94]. However, graphene sheets are prone to stack and agglomerate, resulting in low surface area and uncontrollable pore size. This decreases the electrical double layer capacitance of the electrodes and reduces the CDI performance [95]. Zhang et al. fabricated carbon nanotube–graphene hybrid aerogels as electrode materials through the CO2 supercritical drying of graphene–CNT hybrid hydrogel [96]. This novel material has the impressive properties of a large surface area (435 m2 g−1), large pore volume (2.58 m3 g−1), high conductivity (7.5 S m−1), and being light weight (41.1 mg cm-3). The hierarchically porous structure was found to be crucial to the electro-binding of four different HMs with adsorption capacities of 104.9 mg g−1 for Pb2+, 93.3 mg g−1 for Hg2+, 64 mg g−1 for Ag+, and 33.8 mg g−1 for Cu2+ [96]. Graphene-oxide-modified carbon felt electrodes have been used to dispose of both low-concentration and high-concentration heavy metal pollution using either direct current (DC) or alternating current (AC) electrodeposition. Graphene oxide (GO) provides a high density of surface oxygen functional groups (C-O, COOH, COH, etc.) to assist in electrodeposition. The electrodeposition method showed a capacity two orders of magnitude of a higher capacity (>29 g heavy metal for 1 g of GO) than traditional adsorption methods [97]. DC electrodeposition with GO-modified carbon felt electrodes can reduce single-heavy-metal-ion pollution (Cu, Cd, and Pb) and multiple ion mixtures to below safe water drinking water levels. By tuning the AC frequency and voltage, the electrodeposition method can further selectively recover Cu, Cd, and Pb separately, which adds value to the heavy metal removal process [97].
Based on the electroadsorption theory, the Helmholtz electrical double layer model is widely understood [98]. The initial CDI models for EDL structure models are the Helmholtz EDL model and the Gouy–Chapman–Stern (GCS) EDL model. An advanced equilibrium EDL structure model named the modified Donnan (mD) model can account for the macropores’ confined nature in CDI electrodes [99]. In the mD model, micropores are seen to have a strongly overlapped double layer, so the classical Donnan notion of a uniform pore electrostatic potential can be used. In-depth research has found that a high ion adsorption capacity is closely related to the micropores rather than the mesopores [100,101]. Thus, porous carbon is an attractive material for the electrochemical adsorption of HMs. In these micropores, typically, the Debye length λD is of the order of, or larger than, the pore size. In micropores, when the R of the pores is smaller than the Debye length, an overlap of the electrical double layer is found. Hence, the electrical potential makes a distinct jump from a value on the surface outside the particles to another value in the micropores [80,102]. Then, the concentration of counterions is significantly higher than that on the outside surface. Thus, the micropores (<2 nm) are highly effective in achieving high ion adsorption.

3.2. Heteroatom Doped Carbon

Heteroatom doping is the process of substituting some carbon atoms with other atoms in the carbon structure, primarily P block elements [103]. Bringing heteroatoms (e.g., N, B, S, P, etc.) into carbon nanomaterials can alter the electron configurations of the adjacent carbons due to their sizes and electronegativity differences, hence enhancing electrical conductivity and wettability [103,104]. Moreover, this process can induce a huge number of defects, which can create a more accessible surface area and is beneficial for charge accumulation.
A sheet-like N, O-enriched hierarchical porous carbon material was developed by Zhao et al. [105], adopting eggplant as the carbon source and KHCO3 as the activation agent. The largest capacitance was up to 172.5 F g−1 in NaCl solution. Pb2+-and Cd2+-containing wastewater mainly occurred due to a high concentration of N-doped oxygen functional groups on the surface and a 3D porous structure leading to greater mass transport efficiency and electroconductibility [83]. Further, with wastewater containing Pb2+ and Cd2+ at concentrations of 0.5 ppm and 0.1 ppm, removal rates of 99.1% and 97.9%, respectively, were achieved in rechargeable Zn–air battery-coupled CDI devices. The self-powered system displayed excellent charge efficiency (60.3%) for CDI, whereby the residual concentrations of Pb2+ and Cd2+ in water dropped to 4.5 ppb and 2.2 ppb in 90 minutes. In another study, 3D honeycomb-like porous carbon was synthesized via the hydrothermal carbonization of corncob combined with KOH activation [106]. The newly prepared porous carbon-based electrodes exhibited a high capacitance of 452 F g−1 in 1 M Na2SO4. A batch-mode CDI experiment demonstrated the removal of Cr6+ through a non-faradic process, achieving a removal rate of 91.85% when the initial Cr6+ concentration was 30 mg·L−1. Moreover, a highly porous N-doped graphene-based CDI device exhibited high removal efficiency (90–100%), fast removal (<30 min), and good regeneration performance (10 cycles, 99% retention) for multiple HMs (Pb2+, Cd2+, Cu2+, Fe2+, etc.) in water at a wide range of concentrations (0.05–200 ppm) [94]. Chen et al. developed an N-doped graphene nanosheet material via the one-step pyrolysis of a graphene oxide and cyanamide mixture [94]. Mamaril et al. developed an N-doped and fluorine co-doped graphene oxide as an electrode material, and a specific capacitance of 245.6 F g−1 was achieved with a high removal percentage of 95% and a specific electrosorption capacity of 52.4 mg g−1 at 100 mg L−1 Cu(II) [107]. It is worth noting that highly prepared porous carbon was firstly used as an electrode for membrane CDI to remove multiple HMs, e.g., Pb2+, Cd2+, Fe2+, and Cu2+, simultaneously, as the prepared carbon material showed a high specific surface area of 695 m2 g−1. At 200 ppm, the adsorption capacity of NG for Pb2+ and Cd2+ can, dramatically, reach 521 mg g−1 and 498 mg g−1, and high removal efficiencies of 96%, 80%, 100%, and 100% for Fe2+, Cd2+, Pb2+, and Cu2+, respectively, have been achieved with a wide range of initial concentrations (0.05–200 ppm).
Wang et al. developed a three-dimensional reduced graphene oxide/nitrogen-doped carbon quantum dot composite. In this work, the electrosorption capacity of an rGO/25% NCDs electrode was 19.26 mg/g when the initial concentration of Pb2+ was 5 mg/L [108]. Hu et al. developed two kinds of N-doped carbon with distinct configurations: pyrrolic N-dominated graphene (N-5-G) and pyridinic N-dominated graphene (N-6-G) [109]. The specific capacitance of N-6-G is 239.7 F g−1, which is larger than N-5-G (195.2 F g−1) in Na2SO4 solution, while the Pb2+ electrosorption capacity of N-5-G is 2.43 mg m−2, which is three times that of N-6-G (0.73 mg m−2). The removal efficiency of Pb2+ was found to be 95.8% for N-5-G. Three main types of nitrogen defects exist in the N-doped carbon structure: graphitic nitrogen, pyridinic nitrogen, and pyrrolic nitrogen. The final two kinds are proven to be more electroactive for CDI than graphitic nitrogen because they are prone to donor electrons [109]. Graphitic nitrogen and pyridine nitrogen oxide have shown merit in capacitance. Hu et al. revealed that pyridinic nitrogen groups tend to prefer H+ and Na+, while pyrrolic nitrogen exhibits efficient Pb2+ sorption. This conclusion bridges the matric structure of N-doped carbon with CDI performance for the removal of heavy metal ions.
P and N have the same number of valence electrons, meaning that P-doped carbon materials are also electron-rich. Moreover, the diameter of P is much larger than that of C; thus, P-doping results in the greater local structural distortion of the hexagonal carbon framework and also protrudes out of the graphene plane. The reasons why P-doped carbon materials overcome the steric hindrance effects encountered in N-doped carbon materials were indicated in [110]. Interestingly, multi-heteroatom co-doped carbon materials can produce double functional groups or diatomic synergies, likely resulting in improved electrical conductivity and the enhanced wettability of the carbon materials. The doping of nitrogen atoms into graphitic networks is seen as one of the best approaches to producing n-type conductive materials with improved conductivity [109].

3.3. Polymer Functionalized Carbon Materials

The selectivity of materials for electrochemical adsorption has been highlighted because high selective adsorption can simultaneously achieve the separation and recovery of valuable heavy metals. Despite their large number of functionalities, the inhomogeneous surface properties of carbon materials mean that they mostly exhibit poor selectivity. Functionalization of the carbon surface with designed functional groups not only provides high-density homogenous functional groups but can control the selective chelation of functional groups with designed heavy metal species. Ethylene diamine tetraacetic acid (EDTA) and 3-aminopropyltriethoxysilane functionalized 3D graphene were applied to selectively remove Na(I) and Pb(II) [111]. Interestingly, Na ions were electrically adopted on the 3D graphene pores by electrosorption, while Pb2+ ions were adsorbed on a EDTA-grafted macroscopic 3D graphene cathode (3DEGR) via a chelation reaction. At pH 6.0, using synthetic wastewater containing 20 ppm Pb2+ and 100 ppm Na+, the removal efficiency for Pb2+ was 99.9% and the desorption rate was ~99.6%. The removal efficiency remained at 94.4% after eight cycles, demonstrating the high reversibility of these materials. Thus, heavy metal and salt ions could be separated and recovered, purifying wastewater [111]. The adsorption/desorption experiments indicated that the specific adsorption capacity of the composite for Cu2+ was 99.67 mg g−1. Moreover, polypyrrole/chitosan/CNT (Py/CS/CNT) nanomaterials were prepared via in situ polymerization [112]. The modified nanomaterials provided sufficient porosity for the diffusion of ions and solvent molecules. In the CuSO4 solution of 100 mg L−1, a higher Cu2+ removal rate of 80% was achieved compared with Py/CS-coated electrodes (56.4%). The decreased capacitance of Py/CS/CNT was only 3.4% after 50 cycles, showing remarkable stability due to the highly conductive CNTs which provided an electron pathway and stabilized the polypyrrole and chitosan. Shi et al. used the air-plasma method to modify the structure and composition of polytetrafluoroethylene/carbon nanotube composite materials [113]. Plasma activation raised the relative O and N content, which improved the Pb2+ removal capacity from 1.36 mg g−1 to 2.40 mg g−1 at a voltage of 450 mV and initial concentration of 5 mg L−1.

3.4. Metal Oxide Supported Carbon Composites

Many solutions have been developed to solve the aforementioned challenges. For example, one of the best of such solutions involves using faradaic electrode materials as the electrode. Unlike non-faradaic materials such as porous carbon electrodes, which store salt ions capacitively in an electric double layer along surfaces or nanopores, faradaic electrodes store ions via intercalation or conversion reactions [114]. Incorporated transited metal oxide could efficiently enhance the capacitance and electrosorption ability of carbon-based materials due to the conversion of heavy metal ions and the adsorption of HMs on defects as well as the intercalation of metal ions into the metal oxide layer structure. Banat et al. synthesized highly mesoporous peanut-shell-derived activated carbon/Fe3O4 nanocomposite materials [115]. Cr(VI) was reduced to Cr(III) by Fe(II) ions of Fe3O4 and was then directly electro-captured on carbon-based composite materials in acidic conditions. The synergetic effect of both physical adsorption and electrosorption combined with a reduction pathway was crucial to Cr(VI) removal. Nearly 99.9% of the Cr(VI) was removed with a high electroadsorption rate up to 24.5 mg g−1 at pH 6.5 with an initial concentration of 50 mg L−1. Wang et al. induced highly distributed α-Fe3O4 of a uniform size on carbon matrixes through the carbonization of MIL101(Fe) [116]. The porous carbon yield exhibited surface areas of ~1832 m2 g−1 and showed excellent adsorption capacity for Pb2+ (830.17 mg g−1) and a high removal efficiency of 90.3% at an initial concentration of 500 mg L−1 Pb(II)-containing water through the CDI process with high durability within 10 cycles [116]. This high adsorption capacity and removal efficiency resulted from the electroadsorption and electrodeposition of the Pt(II) on the electrode [116]. In another study, a graphene sheet exfoliated by potassium ions from graphite was prepared [117]. Fe3O4 nanoparticles were successfully supported on porous graphene sheets via the hydrothermal process. The Fe3O4/porous graphene nanocomposites showed great potential for the removal of both organic pollutants and heavy metal ions, with the removal rates for Pb2+, Cu2+, and Cd2+ found to be 90.6%, 92%, and 90.2%, respectively. Moreover, the results revealed ultrahigh electrosorption capacities for Pb2+, Cu2+, and Cd2+ ions which were found to be 47, 40, and 49 mg g−1, respectively, in NaCl solution with an initial concentration of 500 mg L−1, a cell potential of 1.6 V, and a 0.03 mM initial metal ion concentration. α-MnO2 nanoneedles were grown on carbon fiber paper using an ordinary hydrothermal method [118]. Both Ni(II) and Mn(II) ions reduced easily and electrosorbed on the α-MnO2–carbon-paper-based electrode due to its mesoporous structure, contributing to fast electron charge transfer. Birnessite-type MnO2 on carbon nanotubes can effectively remove both Zn2+ and Ni2+ via constant potential electrolysis. CNTs offer a larger specific surface area and higher conductivity to birnessite. Their removal capacities for Zn2+ and Ni2+ have been found to reach 155.6 and 158.4 mg g−1 in 50 mg L−1 origin wastewater, respectively. Hexagonal birnessite is proposed to consist of edge-sharing Mn(IV)O6 octahedral layers, and these layers are interlayered with water molecules. There are Mn(IV) vacancies in Mn(IV)O6 octahedral layers due to the repulsion between neighboring Mn(IV). Heavy metal ions can be adsorbed on or below the vacancies during the isothermal adsorption process [119]. Figure 4 depicts the mechanisms of metal-oxide-supported carbon composites in heavy metal removal. GO–TiO2 nanotube composites have high specific electrosorption capacities of 253.25 mg g−1 and 241.65 mg g−1 for Cu2+ and Pb2+, respectively [120]. This microwave-assisted fabricated composite releases little during desorption at zero voltage caused by induced oxygen, and π electrons lead to strong interactions with Pb2+ and Cu2+. Metal oxide/carbonaceous gel hybrid materials derived from natural and renewable biomass exhibit excellent flexibility and high electrochemical reactivity (capacitance of 120.4 F g−1 in 50 mg L−1 CuCl2 solution), primarily in the area of CDI decontamination applications. In one study, Cu2+ was nearly completely removed (removal capacity of 57.3 mg g−1) and the as-assembled CDI testing apparatus showed durability after 1000 cycles [121]. Activated carbon cloth coated with zinc oxide nanoparticles has been used for electrode inflow via a CDI system. Upon incorporation with ZnO NPs, electrosorption efficiency was enhanced from 17% to 33% for Pb2+, from 21% to 29% for Cd2+, and from 21% to 35% for mixed Pb2+ and Cd2+ ions [122]. These results together illustrate the promising applications of metal-oxide/carbon electrodes for heavy metal removal.

3.5. Others

Aside from carbon-based materials, polymer-based materials have attracted great interest owing to their ion-exchanging ability and conductivity. Polypyrrole is one of the most studied polymers due to its high specific capacitance, high conductivity, and simple preparation. Polyaniline/attapulgite composites (PANI/ATP) were prepared via in situ oxidative polymerization [123]. The calculated equilibrium adsorption capacity was 15.42 mg g−1. The facile modified paper electrode with polymer attached to the surface of ATP contributed to the high removal efficiency of lead ions (66.5%) in wastewater, with an initial concentration of 2.5 mM Pb2+ and stable cyclability after 10 runs (64.2%). Li et al. deposited polypyrrole on chitosan, a stable adhesive agent functionalized with -OH and -NH2 groups, and proved that a polypyrrole/chitosan composite (Py/CS) exhibits considerably stable electroadsorption and desorption behavior and shows good performance regarding the removal of heavy metal ions including Ag+, Cu2+, Pb2+, and Cd2+ [124]. Transition metal sulfide has been viewed as a promising modified candidate for CDI. Hu et al. decorated ZnS and FeS on a carbon felt electrode. The ZnS on the carbon material increased the wettability and electrochemical capacitance. The enhanced adsorption ability of Cu2+ and removal capacity of Cu2+ was 27.4 mg g−1 in the presence of Cr6+ and natural organic matter [124].
The materials used for heavy metal electrosorption are primarily carbon-based materials due to their tunable properties such as surface area, pore size, and distribution as well as their surface functionalities. Overall, polymer carbon composites have been investigated for Cu and Pb removal, and their physical and chemical adsorption abilities were found to be excellent owing to the chelation process of capturing ions. However, a few unusual properties, such as restricted specific area and relatively poor conductivity, limit their performance regarding the electrosorption of HMs. Their adsorption capacity is relatively low, and a higher adsorption capacity was achieved for Pb using PTFE/CNTs. An adsorption capacity of 16.83 mg g−1 for Cu was achieved using polypyrrole/chitosan/CNT due to the use of functional polymer-like polypyrrole, a common conducting polymer, which showed ultra-high specific capacitance and conductivity [112,113]. Moreover, synergy between polypyrrole and chitosan increased adsorption capacity through the chelation process [125,126]. Metal-oxide carbon composites have been proven to show excellent performance in the removal of Pb, Cu, and Cr. Fe3O4/peanut-shell-derived activated carbon material exhibited a 99.9% removal rate and 24.5 mg g−1 Cr6+ adsorption rate in wastewater [116]. The redox ability of transition metal oxide was useful in deducing the toxicity of HMs during the electrosorption process. The removal of Cu by MnO2/carbon fiber demonstrated a 96% removal rate with a 172.9 mg g−1 adsorption rate [127]. The enhancement of adsorption via the polarization potential and unique nanoflower microstructure of deposited MnO2 may equally lead to the superior adsorption ability of MnO2/CF electrodes. TiO2-nanotube-coated CNT electrodes have been proven to efficiently remove Cu and Pb with adsorption capacities of 282.25 mg g−1 and 241.65 mg g−1, respectively. The excellent activity of metal oxide/CNT composite relies on its mesoporous features, good ion-exchanging properties, and very high conductivity. The removal rates and adsorption capacities for Pb by Fe2O3/C were 90.3% and 830.17 mg g−1, respectively; this improvement was accredited to its considerably large specific surface area (1832 m2 g−1) and high dispersion at the near quantum scale. Carbon has been shown to be an excellent material for various HMs, including Pb, Cu, Cr, Cd, and Ni. Carbon is the most widely applied material due to its abundance in nature and facile modification methods. Table 2 presents the various materials that have been studied for heavy metal removal, such as activated carbons, carbon cloths/fiber, ordered mesoporous carbons, carbon aerogels, carbide-derived carbons, carbon nanotubes, carbon black, and graphene.

4. Challenges and Opportunities

Heavy metal removal using CDI techniques provides an energy-efficient and cheap removal method. This real application can be integrated into renewable energy systems, which would be useful in remote areas. Studying the literature on the assessment of CDI energy efficiency reveals its relatively low, specific energy consumption, which is considerably below 1 kWh m−3. Theoretically, energy is stored in the electrical double layer during the charging process, whereby it may be discharged and recovered. However, practically, resistance during charge and discharge leads to energy losses. This resistance arises from the contact resistance of components, connecting wires, the electronic resistance in the solid matrix of the electrode, and ionic transportation resistance. Hence, this requires the optimization of electrode material composition, structure, and porosity as well as minimized contact resistance in the whole CDI assembly.
The pursuit of intercalation materials as promising CDI electrode materials is primarily driven by their ability to yield far higher specific adsorption capacities than conventional carbon-based electrodes. This is due to the ability of intercalation materials to store ions in their solid phase, in contrast to carbon electrodes that store ions by forming EDLs in their micropores. In theory, intercalation materials do not possess a direct advantage over carbonaceous materials considering the reduction in cell resistance and resistive energy loss. Most intercalation materials have poor electronic conductivity, which is inimical to high energy efficiency. However, these materials have two major advantages that allow im-CDI systems to operate in ways that significantly benefit TEE. The first advantage is their high specific adsorption capacity. Although specific adsorption capacity is not directly relevant to TEE, the high SAC of intercalation materials enables im-CDI to remove a considerable amount of salt and thereby achieve an appreciable salinity percentage reduction when employed to desalinate high-salinity feed solutions.
CDI requires an electrode with a large ion-accessible specific surface area. Metal ion electrosorption is a surface-related process, and surface area plays a very important role. However, the entire surface may not have access to metal ions; the latest research shows that not all BET surface areas can participate in ion adsorption. Van der Waals results have revealed that the effectively used are of a porous activated carbon electrode is ∼55% of the BET area of the electrode. The wettability of the electrode and the porous structure of the materials ensure a greater surface area is engaged in ion charge and discharge. Notably, the specific area is highly dependent on the number of small pores, while more small pores limit ion transportation. The final and ultimate limitation on pore size is the metal ion size. Moreover, large pores provide better transportation pathways but limit surface area, while small pores provide a large surface area and higher adsorption capacity but are constrained by ion transportation.
The relationship between these two factors raises vital questions regarding the validity of the gas adsorption models usually used to determine porous electrode surface areas and pore size distributions as well as the effective sizes of ions in pores concerning their degree of solvation. For instance, it is important to discover to what extent and at which potentials the solvation shells of ions entering into pores are partially or completely removed. For example, Li et al. reported an order of Fe3+ (0.62 mg FeCl3/g) > Ca2+ (0.55 mg CaCl2/g) > Mg2+ (0.52 mg MgCl2/g) > Na+ (0.45 mg NaCl/g) for mono-component solutions, while the order became Na+ > Mg2+ > Ca2+ > Fe3+ in mmol/g [137]. Moreover, the removal of certain HMs by CDI was complicated, and an observed order of Cr2+ > Pb2+ > Cd2+ in the percentage removal by Huang et al. was observed under the effect of physical adsorption, which introduced uncertainties into the quantification of the sole electrosorption amount [133]. Thus, a good CDI electrode should not only exhibit good porosity and connectivity; the porosity of the electrode material should also allow for a reasonable combination with macropores, mesopores, and micropores to balance the adsorption surface area and metal ion transportation. Further, dead pores that are not accessible to the ions should be avoided.
The good electrochemical stability of a material over wide pH and voltage ranges plays a key role in real applications. The oxidation of electrode materials under electrochemical conditions is inimical to their stability. The stability of an electrode is also closely related to its high electric conductivity and contact resistance between the porous electrode and the current collector. Higher resistance leads to the generation of local heat, and thus speeds up the corrosion of carbon and the dissolution of metal species. Therefore, in addition to a low-cost design and simple scalability, the design of compacted electrodes with good processability, such as being shapeable into film electrodes based on compacted powders, fibers, or monolith film electrodes, is crucial to improving electrochemical stability.

Author Contributions

Y.X.: writing—original draft and conceptualization; Z.Z.: data curation; X.Z.: validation and visualization; Y.Z.: project administration; W.D.: supervision; Y.C.: writing—review and editing. All authors have read and agreed to the published version of the manuscript.

Funding

This research was funded by the Innovative Province Construction Special Project (2019SK2281), the National Natural Science Foundation of China (No. 42177231), and the Key Research and Development Project of Hunan Province (2019SK2291).

Institutional Review Board Statement

Not applicable.

Informed Consent Statement

Not applicable.

Data Availability Statement

Data sharing is not applicable to this article.

Acknowledgments

This work was financially supported by the Innovative Province Construction Special Project (2019SK2281), the National Natural Science Foundation of China (No. 42177231), and the Key Research and Development Project of Hunan Province (2019SK2291).

Conflicts of Interest

The authors declare no conflict of interest.

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Applsci 13 05635 g001 550
Figure 1. Electrochemical methods for heavy metal removal.
Figure 1. Electrochemical methods for heavy metal removal.
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Figure 2. Summary of the mechanism of CDI for heavy metal removal.
Figure 2. Summary of the mechanism of CDI for heavy metal removal.
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Figure 3. Summary of materials applied for CDI.
Figure 3. Summary of materials applied for CDI.
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Figure 4. Mechanism of metal-oxide-supported carbon composites for heavy metal removal.
Figure 4. Mechanism of metal-oxide-supported carbon composites for heavy metal removal.
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Table
Table 1. Electrochemical methods for heavy metal removal.
Table 1. Electrochemical methods for heavy metal removal.
MetalMethodCathodeAnodeRef
Cr6+Reverse electrodialysisCarbon feltTi/IrO2-Ta2O5[44]
Cr6+Reverse electrodialysisCarbon feltBiotic anode[45]
Cr6+ElectrocoagulationCylindrical vessel iron cathodeVertical rotating impeller Fe anode[15]
Cr6+ElectrocoagulationAir cathodePure sacrificial iron plate[16]
Cr6+Microbial fuel cellCarbon brushCarbon felt[53]
Cr6+ [54]
Cr6+Microbial fuel cellGraphite rodsPolyaniline-modified carbon Fe[55]
Cr6+Alkaline ethanol fuel cellCarbon fiber clothCarbon-fiber-cloth-coated Pt[56]
Pb2+Fuel cellCarbon clothCarbon-cloth-coated Ni[57]
Cr6+Solar electrocoagulationPerforated ZnPerforated Zn[17]
Cu2+; Cr6+; Ni2+;Electrocoagulation–electroflotationStainless steelAl[58]
Cd2+ElectrocoagulationZnZn[18]
Pb2+; Zn2+; Cd2+;ElectrocoagulationStainless steelAl[19]
As3+ElectrocoagulationFe plateAl plate[20]
Cd2+; Cu2+ElectrocoagulationAl plateAl plate[21]
Zn2+ElectrocoagulationStainless steel plateTi expanded mesh[59]
Cu2+Electro-Fenton methodGraphiteFe[32]
Cr3+; Cu2+; Ni2+; Zn2+ElectroflotationStainless steelTi plate[60]
Fe3+; Cu2+; Zn2+ElectrocoagulationIron (carbon steel) platesIron (carbon steel) plates[61]
Cu2+; Zn2+; Ni2+; Ag+Microbial fuel cellActivated carbonCarbon fiber veil[49]
Ni2+; Cu2+Microfluidic electrodialysisAuAu[62]
Mo4+Electro-reductionCarbon fiber clothPt-coated Ti panel[63]
Cd2+; Cu2+; Zn2+; Ni2+; Fe2+; Pb2+; As3+Electro-oxidationTi/RuO2–IrO2Ti/RuO2–IrO2[64]
Hg2+Electro-reduction and ion exchangeGraphite particleTi substrate[65]
Cd2+Electro-reductionPSACGraphite felt[66]
Cu2+Microbial electrolysis cellsStainless steel meshGraphite fiber brush[67]
Cu2+ElectrodepositionTi [68]
Pd2+; Zn2+ Cd2+; Cr6+ Cu2+; Ni2+ Hg2+ElectrodepositionTi [33]
Cu2+ElectrocoagulationConductive carbon clothPt-coated panel[69]
Cu2+; Ni2+ Fe2+ElectrolysisMesh Ti stainless steel plateMesh Pt-coated titanium[70]
Cu2+; Zn2+ Pb2+Microbial electrolysis cellCarbon clothGraphite brush[71]
Ni2+ElectrolysisGraphite plateGraphite felt[72]
Cu2+ElectrolysisGranular graphite [73]
Pd2+; Ni2+ElectrolysisCu sheetPt wires[74]
Cu2+; Cr6+; Ni2+ElectrodialysisPtPt[75]
Cu2+ElectrolysisStainless steelTi grid[76]
Table
Table 2. CDI methods for heavy metal removal.
Table 2. CDI methods for heavy metal removal.
MetalWorking ElectrodeCounter ElectrodeRemoval Rate/Adsorption Rate/CFRef
Cr6+Fe3O4/peanut-shell-derived activated carbonPSAC99.9%;24.5 mg/g; 610 F/g[116]
Pd2+, Cd2+Eggplant-derived carbonZn foil99.1% Pd; 97.9% Cd; 172.5 F/g;[105]
Pb2+Fe2O3@CFe2O3@C90.3%; 830.17 mg/g; 115.6 F/g[116]
Cu2+Polypyrrole/chitosan 56.4%; 99.67 mg/g; 102.96 F/g[124]
Cu2+Polypyrrole /chitosan /CNTGraphite80.08%; 16.83 mg/g; 103.19 F/g[112]
Cu2+ZnS-decorated carbon felt >90%; 27.4 mg/g; 168 F/g[128]
Cr6+Bael-fruit-shell-derived carbon on Cu wireFe wire100%; -; -[129]
Pd2+Graphite foil 90%; -; -[130]
Cu2+Activated carbon on Ti net 95.7%; -; -[131]
Pd2+, Cd2+rGo-CNT on Ti platcocoePt plate/sheet253.25 mg/g Cu, 241.65 mg/g Pb; 165.22 F/g[120]
CrMIL-53(Fe) on graphite paperGraphite paper72.2%; 224.7 F/g[132]
Cr6+Corncob-derived 3D porous carbon on Ni foamPt foil91.85%; 452 F/g;[106]
Pb2+, Cd2+, Cu2+, Fe2+N-doped graphene on graphite foil paperGraphite rod90~96%; 521 mg/g Pb; 498 mg/g Cu[94]
Zn2+, Ni2+Na4Mn14O27 on carbon fabricCarbon fabric90% Zn; 88% Ni155.6 mg/g Zn; 158.4 mg/g Ni[119]
Pb2+, Cu2+; Cd2+Fe3O4/graphene 90.6% Pb; 92% Cu; 90.2% Cd[117]
Cu2+TiO2/watermelon derived carbonaceous hydrogels 57.13 mg/g; 120.4 F/g;[121]
Cd2+; Pd2+; Cr6+Activated carbon cloth on stainless meshActivated carbon cloth on stainless mesh81% Pb; 78% Cr; 42% Cd[133]
Pb2+3D macroscopic graphene on graphite paper3D N-doped macroscopic graphene on graphite paper99.6%; 134.4 F/g[111]
Mn2+Graphene-chitosan-Mn3O4Pt foil12.7 mg/g; 190 F/g[134]
Cu2+Activated carbonActivated carbon69%; 24.57 mg/g; 164.3 F/g[89]
Pb2+, Hg2+, Ag+, Cu2+Graphene/MWCNT hybrid aerogels on Ni foamGraphene/MWCNT hybrid aerogels on Ni foam104.9 mg/g Pb; 93.3 mg/g Hg; 64 mg/g Ag; 33.8 mg/g Cu[96]
Ni2+, Mn2+α-MnO2/carbon fiber paperActive carbon88.9% 16.4 mg/g Ni; 98.5% Mn; 126.7 F/g[118]
Cu2+Carbon aerogelCarbon aerogel73.6%; 29.7 mg/g;156 F/g[92]
Pb2+N-doped grapheneCarbon cloth95.8%; 2.43 mg/m2; 195.2 F/g[109]
Pb2+Chicken-feather-derived carbonChicken-feather-derived carbon81%; 4.1 mg/g[135]
Co2+Activated carbon clothActivated carbon cloth36.54%; 8.3 mg/g[91]
Cu2+Activated carbon fibers 108.8 mg/g Cu; 122.6 mg/g Zn[87]
Pb2+PTFE/CNTs 2.40 mg/g; 43 F/g[113]
Pb2+Polyaniline/attapulgite 66.5%; 12.39 mg/g[123]
Cu2+MnOxSaturated calomelelectrode66%; 372.3 mg/g (electro sob) 44.3 mg/g (iso sorb)[136]
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Xu, Y.; Zhong, Z.; Zeng, X.; Zhao, Y.; Deng, W.; Chen, Y. Novel Materials for Heavy Metal Removal in Capacitive Deionization. Appl. Sci. 2023, 13, 5635. https://doi.org/10.3390/app13095635

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Xu Y, Zhong Z, Zeng X, Zhao Y, Deng W, Chen Y. Novel Materials for Heavy Metal Removal in Capacitive Deionization. Applied Sciences. 2023; 13(9):5635. https://doi.org/10.3390/app13095635

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Xu, Youze, Zhenyu Zhong, Xianhui Zeng, Yuanyuan Zhao, Wenting Deng, and Yuehui Chen. 2023. "Novel Materials for Heavy Metal Removal in Capacitive Deionization" Applied Sciences 13, no. 9: 5635. https://doi.org/10.3390/app13095635

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