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. Author manuscript; available in PMC: 2026 Jun 15.
Published in final edited form as: J Hazard Mater. 2025 Feb 27;490:137735. doi: 10.1016/j.jhazmat.2025.137735

Simultaneous removal of Cr(VI), As(III), and sulfanilamide via an e-barrier electrochemical system: A pilot study

Jong-Gook Kim a,b, Muhammad Fahad Ehsan a, Akram N Alshawabkeh a, Kitae Baek b,*
PMCID: PMC11999765  NIHMSID: NIHMS2062379  PMID: 40024117

Abstract

In-situ electrochemical remediation has emerged as a promising groundwater remediation technology. However, its application has been limited to short-term decontamination. Here, we propose an electrochemical system that combines an e-barrier with pyrite, a sulfide mineral capable of completely removing As(III), Cr(VI), and sulfanilamide continuously for one year. We evaluated the sandbox comprising an e-barrier and pyrite as a flow-through electrochemical reactor on two different scales: (1) a lab-scale small sandbox with sulfanilamide as a model contaminant to assess decontamination performance, and (2) a pilot-scale large sandbox designed for the simultaneous removal of As(III), Cr(VI), and sulfanilamide. The small sandbox achieved 100% removal of sulfanilamide, demonstrating the effectiveness of the combined system. The large sandbox demonstrated 100% removal efficiency against contaminants mixture for up to one year, with effluent maintaining a neutral pH even without an external neutralizing process. This remarkable performance was attributed to the activation of pyrite by anodic oxygen (O2), producing dissolved iron that leads to the formation of iron hydroxide (e.g., green rust), which serves both as an adsorbent and precipitant for contaminants. Our findings indicated that the combination of electrochemical reactions and pyrite is an effective approach for the simultaneous removal of organic and inorganic contaminants.

Keywords: Groundwater remediation, electrochemical oxidation, pyrite, e-barrier, pilot-scale

1. Introduction

Groundwater contamination is a pressing global concern that significantly impacts human health and ecological systems (Li et al., 2021; Zhang et al., 2022). Several in-situ and ex-situ remediation approaches have been applied to groundwater remediation (Hussain et al., 2022). However, chemical oxidation has received attention due to its great efficiency and fast reaction rate (Baciocchi et al., 2014). Several oxidants, such as ozone (O3), hydrogen peroxide (H2O2), and permanganate (MnO4) were generally applied to carry out the oxidation of target contaminants in groundwater. However, they have a short lifetime and continuous injection is required to achieve complete removal (Huling and Pivetz, 2006; Khan and Carroll, 2020; Liang et al., 2014). Additionally, most oxidants injected into the subsurface are rapidly consumed through side reactions with natural organic matter and materials, resulting in significant oxidant loss and reduced efficiency. In contrast, electrochemical oxidation (EO) has emerged as an alternative technique to overcome conventional oxidation due to the direct in-situ generation of reactive species at the electrode surface, minimizing losses and side reactions (Dai et al., 2016; Kim et al., 2023; Zhang et al., 2019).

Despite the excellent efficiency, EO encounters some challenges in a groundwater environment such as the lack of a pilot-scale or field applications to provide field applicability and design parameters (Monteil et al., 2021; Thiam et al., 2018). Most studies reported up to date suggested the complete removal of just targeting contaminants at the pilot scale without considering the groundwater environment (Monteil et al., 2021; Tawabini et al., 2020; Zhu et al., 2010). Accordingly, these results and parameters including short duration, small volume, and presence of various minerals are not practical and available to simulate the groundwater conditions (Benekos et al., 2021; Hyldegaard et al., 2019). In this context, it is necessary to consider the presence of sulfide minerals when simulating groundwater environments. These minerals are commonly found in the subsurface, where they may interfere with groundwater remediation. They are widely distributed with thermodynamically stable forms, particularly pyrite (FeS2) (Stolze et al., 2022; Zhu et al., 2022). Groundwater remediation can be problematic due to the dissolution and accumulation of heavy metals during weathering and exposure to oxygen (Wu et al., 2020). That is, the sulfide mineral can lead to potential deterioration such as acid mine drainage (AMD). Despite this problem, few studies have been explored at a pilot-scale with EO in combination with sulfide minerals. As a response to this challenge, this study proposed the field applicability and design parameters of the pilot-scale with EO in the presence of sulfide minerals.

Another challenge is the mixed contamination of organics and potentially toxic elements (PTEs) (such as As(III), Cr(VI), and Se(VI)) in groundwater (Papazotos, 2021; Tomašek et al., 2022). Even though EO has been successfully applied to mineralize or detoxify various organic pollutants, only electrokinetic remediation has been applied to transport or separate PTEs in the unsaturated zone. Some redox-sensitive PTEs can be transformed electrochemically into less toxic or less mobile species (As(V), Cr(III), and Se(IV)). There have been few studies that have attempted the complete removal of PTEs and the simultaneous removal of organics and PTEs by electrochemical reactions. To address this problem, the physical/chemical addition or composite development can be applied. Electro-Fenton using Pd/Fe3O4 nanocatalyst degraded simultaneously Cr(VI) and humic acids(Huang et al., 2017). Additionally, the simultaneous oxidation and adsorption of As(III) was proceeded with Fe/Fe3C@CNTs (Wang et al., 2021). Considering the co-presence of organic and inorganic matter in groundwater, it is essential to treat the mixture simultaneously (Qian et al., 2017). Thus, the simultaneous electrochemical removal of organic pollutants and redox-sensitive PTEs can be achieved and should be evaluated properly for the potential in groundwater remediation.

In this context, the present work evaluated the scale-up and electrochemical removal of organic pollutants and PTEs in the simulated groundwater. A previous study suggests that the combination of EO and pyrite can effectively remove contaminants from groundwater (Kim et al., 2021; Kim et al., 2024). Accordingly, the purpose of this study is to demonstrate the practical applicability of a combined system by conducting pilot-scale experiments to determine whether groundwater contaminants could be effectively removed. Sandbox was employed to simulate the subsurface environment in the presence of pyrite. The combined system with electrochemical oxidation and pyrite was assessed to remove the contaminants mixture of As(III), Cr(VI), and sulfanilamide. Additionally, the long-term feasibility of the system was also evaluated for 1 year. A lab-scale small sandbox was used to evaluate the effectiveness of the combined system for just organic pollutants. A mixture of organic and inorganic contaminants was assessed for about a year in a pilot-scale large sandbox.

2. Materials and Methods

2.1. Materials

Sodium arsenite (NaAsO2), sulfanilamide (≥99.0%, H2NC6H4SO2NH2), sodium bicarbonate (99.5%, NaHCO3), potassium chloride (≥99.0%, KCl), and pyrite (iron(II) sulfide, FeS) were purchased from Sigma-Aldrich (Seoul, Korea). Sodium chloride (≥99.0%, NaCl) and anhydrous sodium sulfate (≥99.0%, Na2SO4) were purchased from Daejung Co. (Korea). Anhydrous magnesium sulfate (MgSO4) was purchased from OCI Co. Ltd. (Korea). Pyrite was washed several times with de-ionized water and a hydrochloric acid solution (HCl). Subsequently, pyrite was milled with mortar and sieved under 100 mesh screens to get a homogeneous sample. The washed pyrite was dried and stored in a desiccator to avoid oxidation by contacting moisture and oxygen. For the electrochemical experiments, 0.33 wt% pyrite (mixed with sand) was placed between the anode and cathode. Synthetic groundwater consisted of 8.20 mM NaHCO3, 4.20 mM MgSO4, and 3.44 mM KCl (Kim et al., 2022). Carbon fiber (CF) electrode (Model name: H2550, Density: 1.80 g/cm3, diameter: 7.0 μm) was purchased from HYOSUNG Co., Ltd., Korea.

Among PTEs, As(III) and Cr(VI) were chosen because both are very mobile and toxic, and their redox potential makes them difficult to remove from groundwater (Shankar and Shanker, 2014). For the complete removal of both contaminants by adsorption or precipitation, the oxidation or reduction of As(III)/Cr(VI) to As(V)/Cr(III) was required. In addition, World Health Organization (WHO) guidelines also recommend a limit of 10 μg/L of As and 50 μg/L of Cr(VI) in water (Guan et al., 2011). Despite this, the actual groundwater was contaminated with arsenic and chromium at levels ranging from μg/L to hundreds of mg/L (Lu et al., 2020; Poonia et al., 2021). Thus, 8.5 mg/L As(III) and 6.0 mg/L Cr(VI) were selected as inorganic contaminants. As a model organic contaminant, 0.05 mM sulfanilamide, a sulfur-based antibiotic, was used.

2.2. Sandbox system

A previous study examined different sandbox configurations for sulfanilamide oxidation, including single well, double well, and e-barrier (Kim et al., 2024). The EO system with an e-barrier was the best configuration in terms of overall performance and effluent pH. E-barriers are electrically induced redox barriers for groundwater treatment that are both cost-effective and superior in performance (Gent et al., 2012; Wani et al., 2006). Thus, the e-barrier system was chosen to simulate the groundwater environment. A conceptual diagram of the e-barrier can be seen in Figure 1.

Figure 1.

Figure 1.

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Front and side view of sandbox (a) lab-scale small sandbox, (b) pilot-scale large sandbox

A small 75 L acrylic sandbox with dimensions 0.5 m (L) × 0.5 m (W) × 0.3 m (H) was used for the bench experiments in the lab-scale small sandbox (Fig. 2). For the pilot studies, a 375 L acrylic sandbox with dimensions 1.5 m (L) × 0.5 m (W) × 0.5 m (H) was evaluated in a pilot-scale large sandbox. To simulate the groundwater environment, the silica sand (0.4 – 0.7 mm) was densely packed into the sandboxes. The silica sand sample was washed with a weak acidic reagent (0.1M HCl) for 30 min, and then the water was cleaned to obtain a homogenous sand sample. In the small e-barrier system (Exp. 1), the wells were built on both sides of the sandbox to install electrodes. The 15 cm apart anode and cathode were positioned on the front and the backside of the sandbox. The pyrite, 0.33 wt% of the total sand, was placed between the electrodes. In the pilot scale of the e-barrier, two e-barriers with a set of anode zone – pyrite zone – cathode zone were installed for more continuous and complete removal of contaminants in the system (Fig. 1). In the lab-scale small sandbox, ports 1 and 2 were positioned between the influent and the anode, while ports 3 and 4 were located between the cathode and the effluent (Fig. 2). In the pilot-scale large sandbox, ports 1 and 2 were positioned between the influent and anode, ports 3 and 4 between the cathode and anode, and ports 5 and 6 between the cathode and effluent. Samples were collected from these ports and averaged for concentration calculation. The e-barrier is similar to the permeable reactive barrier (PRB), where the electrodes also function as reactive materials (Fig. 2 and Table 1). Additionally, the bundle of 20 CF wires (7.41 g), while each CF wire is 46 cm long and weighs approximately 0.37 g, was used as an anode while 1/4th of this amount was employed as the cathode to adjust the pH. A current of 15 mA was supplied, and a flow rate of 0.5 mL/min was used.

Figure 2.

Figure 2.

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Schematic diagram in the combined system of Exp. 1 (lab-scale small sandbox) and Exp. 2 (pilot-scale large sandbox)

Table 1.

Detailed experimental conditions to evaluate the e-barrier

Exp. # Process Sulfanilamide (mg/L) As (III) (mg/L) Cr(VI) (mg/L) Pyrite dosage Current (mA) Flow rate (mL/min) Electrolyte (mM)
1 EO + Pyrite 8.61 (0.05 mM) - - 0.33 wt% 15 0.5 Synthetic groundwater
2 8.0 6.5
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2.3. Analytical methods

A 2 mL aqueous sample was periodically collected and filtered through a 0.45 μm syringe filter to remove suspended particles. The total arsenic concentration was measured using inductively coupled plasma optical emission spectrometry (ICP-OES). To differentiate between As(III) and As(V), an anion exchange filter was employed, capturing As(V) while allowing As(III) to pass through (Kim et al., 2024). Cr(VI) concentration was determined through the diphenyl carbazide (DPC) method using a UV-Vis spectrometer, while total chromium and dissolved iron concentrations were analyzed using ICP-OES (Kim et al., 2024). Sulfanilamide content was measured by high-performance liquid chromatography (HPLC) using a C18 column with a mobile phase of methanol and deionized water (Kim et al., 2024). The surface morphology and composition of the samples were examined via scanning electron microscopy coupled with energy dispersive spectroscopy (SEM-EDS). Finally, X-ray diffraction (XRD) was used to characterize the crystalline structure of the materials.

3. Results and Discussion

3.1. Pilot e-barrier system for removal of sulfanilamide

The e-barrier system combined with pyrite was evaluated in terms of sulfanilamide removal and effluent pH for 1 year using the bench-scale small sandbox. This system hypothesizes that the anode zone transforms and oxidizes the sulfanilamide by anodic oxidation and generates oxygen via water electrolysis, which activates pyrite and produces dissolved iron. To some extent, the pyrite also facilitates adsorption of heavy metals. At the cathode zone, even though it does not contribute to direct oxidation, it can neutralize the effluent pH (Wang et al., 2024). At the same time, some iron materials can also be formed resulting from the reaction of dissolved iron with hydroxyl ions from pyrite and cathode, and also acts as a potential remediation agent.

As expected, the anode zone completely removed sulfanilamide (Fig. 3a). Although CF oxidation over time led to a decrease in sulfanilamide removal, it returned immediately after oxidized CF was replaced with a new CF. Simply by reversing the polarity, the e-barrier achieved continuous removal without complicated excavation or installation processes. The anode zone by itself was capable of removing the sulfanilamide without pyrite addition. Concerning pH, the anode zone lowered the pH sharply to 2.0, the pyrite zone maintained the pH, and the cathode zone increased the pH to 8.0, and finally, the effluent pH was neutral at 7.5 (Fig. 3b). As expected, H+ and OH− from the anode and cathode keep the acidic and alkaline conditions (Eq. 1 and 2). Thus, the e-barrier did not require any additional neutralization process.

Anode:2H2O→O2+4e−+4H+ (1)
Cathode:2H2O+2e−→H2+2OH− (2)

Figure 3.

Figure 3.

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Lab-scale small sandbox e-barrier system in Exp. 1 (a) sulfanilamide removal (b) pH change (c) ORP and (d) iron (Fe) dissolution

After the anode zone, the pyrite was activated electrochemically to generate ferrous (Fe2+) by reaction with anodic oxygen. Pyrite was stable originally, however, iron was released by the electrochemical activation process in the pyrite zone (Fig. 3d). The iron concentration in the pyrite zone increased sharply to 150 mg/L, where ferrous was a major species. Later, the iron concentration released in the cathode zone was lowered to < 1 mg/L since ferrous precipitated as hydroxide at the cathode in the form of Fe(OH)2 or Fe(OH)3. As evident from the pH-ORP diagram in Fig. S1, it directly corresponds to the formation of Fe(OH)2 or Fe(OH)3 in the cathode zone. The ferrous hydroxide can be a reducing agent, and Fe(OH)3 after oxidation serves as an adsorbent for heavy metals. The dissolved iron from pyrite may have formerly been the source of the AMD problem. The e-barrier system prevents AMD formation by balancing anodic oxidation and cathodic reduction processes, inherently stabilizing pH and precipitating dissolved iron, eliminating the need for additional neutralization steps. The ferrous was precipitated as a form of iron hydroxide. A pH and ORP of the system endorses the ferrous hydroxide formation. In sections 3.2 & 3, this is discussed more in detail with evidence suggested by XRD.

Based on the results, the e-barrier system effectively oxidized sulfanilamide by anodic oxidation. Additionally, the system showed stable performance for 1 year hence proving the long-term stability of the e-barrier system. Moreover, the e-barrier system discharged neutral effluent thus avoiding any additional neutralization. As expected, the e-barrier system showed good oxidation performance and discharged neutral effluent without additional pH adjustment.

3.2. Pilot e-barrier system for removal of As(III), Cr(VI), and sulfanilamide

A pilot-scale e-barrier system in a large sandbox was operated to treat the mixture of sulfanilamide, As(III), and Cr(VI) simultaneously for 1 year. The sulfanilamide was completely oxidized in the first anode zone (Fig. 4a). The pH decreased sharply to 2.0 in the first anode zone, increased suddenly to 8.5 in the cathode zone, and finally discharged at pH 7.7 in the effluent (Fig. 4b). In the anode zone, the ORP increased to 500 mV indicating oxidative condition in the anode zone, and decreased to – 300 mV in the cathode zone. Finally, the ORP in the effluent was 50 mV (Fig. 4c).

Figure 4.

Figure 4.

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Pilot scale of EO process (a) Sulfanilamide (b) pH and (C) ORP

As(III) and Cr(VI) were successfully removed in the e-barrier system (Fig. 5). As(III) was lowered from 8.0 mg/L to 6.8 mg/L in the anode zone due to the adsorption by sand. Following, As(III) concentration in the pyrite zone decreased from 6.8 mg/L to 0.0 mg/L because of pyrite adsorption and precipitation with As(III)/As(V). In addition, the electrochemical activation of pyrite via anodic oxygen also generates iron oxide, which can contribute to the adsorption and precipitation with As(III)/As(V). In addition to adsorption, As(III) is oxidized to As(V) at the anode during EO, making As(V) less mobile and easier to adsorb and precipitate [34]. Oxidized As(V) was transferred to the pyrite zone, where the adsorption of As(V) at pyrite and the precipitation with dissolved iron and As(V). In the pyrite zone, Fe concentrations were observed to reach up to 200 mg/L, functioning as precursors for the formation of iron oxide. In Fig. 6a, XRD analysis of the minerals from the pyrite zone showed several iron-based minerals (Goethtie and magnetite), the precipitate with iron arsenate, and both As(III) and As(V). Based on the results, it can be concluded that the combined system not only successfully oxidizes As(III) to As(V), but also electrochemically activates the pyrite to produce the dissolved iron. In the end, this process facilitates the complete removal of both As(III) and As(V).

Figure 5.

Figure 5.

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Pilot scale of EO process (a) Arsenic (As) (b) Chromium (Cr) and (c) Iron concentration

Figure 6.

Figure 6.

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XRD analysis of mineral after EO process (a) pyrite zone (b) cathode zone

The reduction of Cr(VI) concentration from 6.5 mg/L to 4.0 mg/L within the pyrite zone is attributed to the electro-generation of iron oxides, which facilitated a slight decrease in Cr(VI). However, a more substantial reduction occurred within the cathode zone, where Cr(VI) was completely eliminated, unlike As(III), which was primarily removed in the pyrite zone. The distinct efficiency of Cr(VI) removal in the cathode zone is due to its strong reducing environment, which not only reduces Cr(VI) but also converts Fe(III) to Fe(II). Fe(II), in turn, plays a critical role in further reducing Cr(VI) to Cr(III), rendering it more amenable to adsorption and precipitation (Joe-Wong et al., 2017). In the cathode zone, iron hydroxide-based compounds such as lepidocrocite, magnetite, and green-rust, along with iron hydroxide minerals, were identified (Fig. 6b). These compounds contribute to the transformation of Cr(VI) to Cr(III). During this process, Cr(III) combines with hydroxide ions (OH−), which are produced through water electrolysis, facilitating the precipitation of Cr(III) as a stable compound. This process effectively transfers the precipitated Cr(III) into the cathode zone, where it is sequestered from the system. The condition of the cathode zone is favorable to form Fe(II) and Cr(III), confirming the reduction of Cr(VI) to Cr(III) in the cathode zone (Fig. S1 & S2). This mechanism ensures the efficient and complete removal of Cr(VI) from the system. Moreover, this integrated system has demonstrated consistent removal of As(III), Cr(VI), and sulfanilamide over an operational period of one year. Further insights into these removal mechanisms are elaborated in Section 3.3.

3.3. A plausible mechanism of the combined system

A plausible mechanism of the combined system was explained in Fig. 7. At the anode zone, the anodic reaction has an oxidative potential to oxidize directly sulfanilamide and As(III). At the same time, the by-product (O2) generated at the anode via water electrolysis was used as a supplementary source to activate the pyrite. At the pyrite zone, the electrochemically activated pyrite produced the dissolved iron. Due to the oxidative environment of the pyrite zone, species such as As(V), O2, and Fe(III) are slightly more prevalent than As(III), OH−, and Fe(II). Consequently, Fe(III)-rich minerals like goethite and magnetite are formed, and As(V) is detected (Fig. 6a). These iron oxides play a critical role in effectively removing both As(III) and As(V) from the pyrite zone. In contrast to As(III), Cr(VI) is not fully removed in the pyrite zone but is completely eliminated in the cathode zone. This difference arises because Cr(VI), under oxidizing conditions, is less readily reduced compared to Cr(III). Although goethite and magnetite in the pyrite zone can adsorb and remove some Cr(VI), the reduction of Cr(VI) to Cr(III) is limited due to the presence of O2. However, in the cathode zone, the dominant species are Fe(II), OH−, and Cr(III), while Fe(III), O2, and Cr(VI) are less prevalent. This creates a favorable environment for the simultaneous reduction and removal of Cr(VI) to Cr(III). Additionally, Fe(II)-based compounds in the cathode zone contribute significantly to the effective removal of Cr(III). Particular importance is the formation of green rust, which is a layered iron hydroxide structure rich in both Fe(II) and Fe(III). Green rust is highly effective in reducing Cr(VI) to Cr(III). The cathode zone provides favorable conditions for the formation of green rust, such as a mildly alkaline pH, a reducing environment, the presence of OH− generated from the cathode, and the availability of various anions (CO32−, SO42−) in the electrolyte. Thus, green rust produced in the cathode zone further facilitates the complete reduction and removal of Cr(VI). In summary, the system’s sequential redox reactions and the formation of various iron oxides and hydroxides enable the effective removal of both organic pollutants and redox-sensitive PTEs. This process ensures the comprehensive treatment of contaminants through the system’s redox transformations and adsorption mechanisms.

Figure 7.

Figure 7.

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(a) XRD analysis (b) SEM image of mineral

XRD and SEM analyses provided additional evidence supporting the electro-generation of various materials (Fig. 7). XRD analysis shows the presence of several iron oxide and hydroxide phases, with distinct peaks observed at 21.3°, 23.1°, 35.4°, and 36.4°, corresponding to the formation of goethite, green rust, magnetite, and lepidocrocite, respectively [36, 37]. Additionally, a broad scattering peak was detected, indicating the presence of amorphous material. Although this amorphous peak is typically associated with green rust, XRD analysis suggested that it comprised green rust, lepidocrocite, goethite, and magnetite, likely due to the unstable and amorphous nature of green rust [38]. These materials, characterized by their Fe(II)/Fe(III) content, belong to the family of layered double hydroxides (LDHs), which are known for their strong pollutant removal capabilities [38].

SEM imaging further supported these findings, revealing several aggregations and hexagonal crystalline structures. The aggregations were indicative of iron hydroxide, while the sharp hexagonal crystals were identified as green rust [39]. Additionally, the iron oxide with its cubic or crystal structure, was also observed, appearing either isolated or in clustered formations. This observation suggests the coexistence of iron hydroxide and green rust on the cathode surface, which is consistent with their roles in pollutant removal [40]. Similar removal mechanisms have been reported in previous studies, indicating that the system’s ability to generate iron oxides/hydroxides and green-rust is crucial for effective contaminant removal. In summary, the combination of XRD and SEM data corroborates the formation of various iron-based materials, including green rust, lepidocrocite, goethite, and magnetite. These compounds, particularly those containing Fe(II) and Fe(III), play a pivotal role in the removal of redox-sensitive contaminants, in line with the previously identified removal mechanisms.

Consequently, the by-product of O2 and OH− from the anode and cathode was used as a supplementary source for the removal of a mixture of organic and inorganic pollutants. The electrochemical reaction and pyrite have synergistic remediation of groundwater of groundwater for the long term.

3.4. Future perspective

The above results show the feasibility of EO in the presence of pyrite to treat groundwater contaminated with a mixture of organic and inorganic pollutants. The e-barrier system operated for 1 year demonstrates good performance during the test period, while also generating neutral effluent at the pilot scale. Considering the long-term stability of the process, the e-barrier system proposed in this study provides operating parameters to design and apply in terms of duration, volume, electrolytes, removal of heavy metals, and effluent pH. In addition, given the recommended effluent pH range (6.5 – 8.5), this process is significantly advantageous and simply requires adjusting the electrode configuration without requiring any further effluent treatment (Rezaei and Hassani, 2018; Tay, 2012). Therefore, based on these findings, future work should focus on real groundwater and cost evaluation.

Furthermore, this system is highly efficient in terms of energy consumption, which is a key consideration for EO applications. Energy consumption was estimated to be 3.97 and 7.30 kWh/m3 for small and large sandboxes, respectively (Table 2), while other studies ranged from 0.79 to 173 kWh/m3 (Table S1). It should be noted that the e-barrier system achieved energy efficiency. Additionally, Table 3 summarizes several studies of electrochemical reactions at a pilot-scale for contaminant removal. While most published literature lacks in simulation of the real groundwater conditions and addresses organic pollutants (e.g. phenol and COD) only, this study assessed the mixture of organic and inorganic pollutants while utilizing simulated groundwater. Not only were the groundwater electrolyte and pH conditions properly modeled, but the evaluation of long-term changes including small reactor volume and short reaction time was also considered. Groundwater remediation should consider both contaminant and site characteristics. As a pilot-scale study, this study provides valuable insights into the feasibility of groundwater remediation based on various characteristics.

Table 2.

Energy consumption by pilot e-barrier experiments

Sample Voltage (V) Current (A) Time (day) Volume (m3) Amount of power supply Energy consumption (kWh/m3)
e-barrier (Exp. 1) 5.67 0.015 365 0.375 2 3.97
e-barrier (Exp. 2) 5.21 4 7.30
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Table 3.

Summary of previous studies in electrochemical oxidation

Experiment Pollutant Process Electrolyte Flow rate (mL/min) Reactor Removal efficiency (%) Initial pH Final pH Reactor volume (L) Duration Current Anode-Cathode Reference
1 Phenol EO - 1.66 L/min Pilot scale 99.9 7.5 ~ 9 7.1 ~ 7.4 22.2 24 min (Abou-Taleb et al., 2021)
2 TCE EO + Biodegradation NaHCO3 15 Pilot scale 73 7.5 ~ 7.7 7.2 ~ 7.4 28.8 50 days (Yuan et al., 2021)
3 Hydrochlorothiazide EO Na2SO4 20.4 Pilot scale 97 3 - 1.89 16.6 mA/cm2 BDD-Carbon felt (Monteil et al., 2021)
4 Phenol EP Na2SO4 56.25 Flow-through 77% - - 4.5 80 min 8.5mA/cm2 BDD- SS (Zhu et al., 2010)
5 1,4-Dioxane + TCE EO + Biodegradation Groundwater 1.07 Flow-through 1,4- Dioxane (88.45%) +TCE (62%) 6.9 4.2 - 1.2 mA/cm2 Ti/IrO2−Ta2O5 (Jasmann et al., 2017)
6 1,4- Dioxane - Groundwater 0.32 L/h Pilot scale 8.0 8.5 month IrO2-Ta2O5 (Blotevogel et al., 2019)
7 TOC EF Urban wastewater 200 L/h Pilot scale 63.5 3 2.5 360 min 50 mA/cm2 RuO2 - Carbon-PTFE air-diffusion (Thiam et al., 2018)
8 BTEX EO Seawater 0.2 m3/h Pilot scale 40.8% benzene, 86.1%toluene 100%, ethylbenzene 100% xylene 9.05 7.37 20 120 min 100 mA/cm2 BDD – Carbon-PTFE (Tawabini et al., 2020)
9 2,4-dichlorophenoxyacetic acid (2,4-D) EO Na2SO4 4 – 10 L/min Pilot scale 70 3 - 3 160 min 31 mA/cm2 BDD – Cathode (Garcia et al., 2013)
10 Methyl orange EO Na2SO4 12 L/min Pilot scale 60.3 3 - 3 138 min 31 mA/cm2 BDD – SS (Ramirez et al., 2013)
11 COD EO Na2SO4 10 L/h Pilot scale 51 - - 500 50 h 10 mA/cm2 BDD – SS (Dominguez-Ramos and Irabien, 2013)
12 COD EO Na2SO4 151 dm3/h Flow-through 93.6% (Ti/Pt) 97.1% (BDD) 7.59 7 – 8.5 5 8 h 40 mA/cm2 Ti/Pt or BDD – Ti disc (dos Santos et al., 2014)
13 Phenol EO Na2SO4 0.4 m3/h Pilot scale 100 7 - 5 8 h 40 mA/cm2 BDD – Carbon-PTFE GDE (Tawabini et al., 2020)
14 Phenol EO Na2SO4 8.26 – 33.06 mL/L Flow-through 75 3 – 11 - 4.5 8 h 15–30 mA/cm2 BDD – SS (Zhu et al., 2011)
15 TOC EO SO42-/Cl− - Pilot scale 90% TOC 100% N-NH3 7.4 – 10.3 - 250 5 h 300–700 mA BDD – SS (Anglada et al., 2010)
16 COD EO Na2SO4 / NaCl 4,500 dm3/h Pilot scale 37.08 – 90.28 Color removal 20.1 – 81.3 COD removal 10.3 – 10.4 - 25 - 35, 60, 100 mA/m2 BDD – SS (Alvarez-Pugliese et al., 2016)
17 Tetrahydrofuran EO Na2SO4 450L/h Pilot scale 98% COD 95% TOC - - 200 5 h 600 A /m2 Si/BDD – SS (Urtiaga et al., 2014)
18 Sulfanilamide, As(III), Cr(VI) EO Synthetic groundwater 0.5 mL/min Pilot scale 100% sulfanilamide, As(III), Cr(VI) 7 – 8 7 – 8 324 1 year 15 mA CF – CF This study
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Electrochemical oxidation: EO; Electro-Fenton: EF; Stainless steel: SS; Carbon fiber: CF;

4. Conclusions

The pilot-scale e-barrier system, integrated with pyrite, successfully treated groundwater contaminated with a mixture of sulfanilamide, As(III), and Cr(VI) over a one-year period. The system demonstrated effective removal of contaminants through a combination of anodic oxidation, pyrite activation, and cathodic reduction. Sulfanilamide was fully oxidized in the anode zone, while As(III) was oxidized to As(V) and adsorbed by iron oxides and hydroxides formed in the pyrite zone. Cr(VI) was reduced to Cr(III) in the cathode zone, where iron-based compounds such as green rust, magnetite, and lepidocrocite played a significant role in adsorption and precipitation. A key feature of the system was its ability to maintain a neutral effluent pH, eliminating the need for additional pH adjustment. The electrochemical activation of pyrite and the formation of iron hydroxides and oxides simultaneously facilitated the removal of heavy metals and organic pollutants. The presence of green rust in the cathode zone enhanced Cr(VI) reduction, providing an efficient pathway for its complete removal. The e-barrier system demonstrated long-term stability and consistent performance, making it a promising technology for groundwater remediation. However, field trials and cost evaluations are necessary to further assess its full-scale applicability. Overall, the study highlights the potential of this innovative system for addressing multi-contaminant groundwater pollution through a sustainable and energy-efficient approach.

Supplementary Material

S.I.

Acknowledgment

Funding:

This study was partially supported by the Superfund Research Program of the National Institute of Environmental Health Sciences (NIEHS), National Institutes of Health (NIH; grant number P42ES017198)

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S.I.
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