L-Cysteine Enhanced Degradation of Chlorobenzene in Water Using Nano Zero-Valent Iron/Persulfate System

Abstract

Nano zero-valent iron (nZVI) particles have received much attention in environmental science and technology due to their unique electronic and chemical properties. While sulfate radical-based advanced oxidation processes (SR-AOPs) activated by nZVI show promise for mono-chlorobenzene (MCB) degradation, their efficiency is severely limited by surface oxidation of nZVI and Fe³⁺ accumulation. This study aims to enhance the nZVI/persulfate (PS) system using L-cysteine (Cys) to achieve effective MCB removal. The work involved synthesizing nZVI via borohydride reduction, followed by comprehensive characterization and batch experiments of the Cys/nZVI/PS degradation system of MCB were carried out to evaluate the key influencing factors and analyze the reaction mechanism of Cys-enhanced MCB degradation. Under optimal conditions (0.1 g/L nZVI, 3 mM PS, 0.1 mM Cys, pH 3), 92.6% of MCB was degraded within 90 min—an 18.7% improvement compared to the Cys-free system. Acidic pH promoted Fe²⁺ release and significantly enhanced degradation, while HCO₃⁻ strongly inhibited the process. Mechanistic studies revealed that sulfate radicals (SO₄^•−) played a dominant role, and Cys served as an electron shuttle that facilitated the Fe³⁺/Fe²⁺ cycle and enhanced Fe⁰ conversion, thereby sustaining PS activation. This study demonstrates that Cys effectively mitigates the limitations of nZVI/PS systems and provides valuable insights for implementing efficient SR-AOPs in treating chlorinated organic contaminants.

1. Introduction

Mono-chlorobenzene (MCB) is an organic compound with strong volatility, toxicity, stable chemical properties, and low biodegradability. It is widely used in manufacturing, agriculture, electronics, textiles, and pharmaceuticals [1]. The volatile toxicity, stability, low solubility, and difficult degradation properties of MCB enable it to persist in the ecological environment for a long time and undergo long-distance migration, resulting in its impact on various parts of the world [2]. MCB pollution is very common in soil and groundwater. For example, the groundwater pollution problem of an air force exercise base in Georgia, the United States, is quite typical. A large number of pollutants, such as chlorobenzene and trichloroethylene, were detected in the groundwater [3]. Megahed et al. detected a large amount of chlorobenzene pollutants in the water of the Nile Delta in Egypt [4]. MCB pollutants can increase the risk of cancer in humans, cause mutations in animal, plant, and human cells, and even lead to deformities in newborns. In the process of creating a harmonious ecology, countries around the world have placed strict control over MCB pollutants in an important position.

Sulfate radical-mediated advanced oxidation processes (SR-AOPs) represent a cutting-edge class of oxidation techniques, leveraging sulfate radicals (SO₄^•−) as their operative transient species to decompose environmental pollutants [5]. Relative to hydroxyl radicals (^•OH), SO₄^•− radicals exhibit greater redox strength (E₀ = 2.5–3.1 V), a more extended existence (half-life: 30–40 μs), and a preferential interaction with contaminants featuring unsaturated or aromatic moieties [6]. A variety of transition metal ions, including Fe²⁺, Co²⁺, Cu²⁺, and Ag⁺, have been explored for triggering the decomposition of persulfate [5], with iron salts gaining widespread adoption due to their favorable balance of affordability, environmental compatibility, and reactivity [7]. Persulfate can be efficiently converted into active SO₄^•− radicals through activation by Fe²⁺ [8]. This Fe²⁺-catalyzed persulfate process demands an activation energy of 12 kcal mol⁻¹, a figure considerably lower than the 33.5 kcal mol⁻¹ threshold associated with the thermal breaking of the O-O bond [9]. The chemical pathways associated with the Fe²⁺/PS mechanism are depicted in Equations (1) and (2).

Conversely, inherent limitations of the Fe²⁺/PS method may hinder its widespread adoption in practical settings. For instance, its efficacy largely relies on acidic environments [8]. Moreover, the generation of SO₄^•− may be diminished as they react with surplus Fe²⁺ under certain conditions, as illustrated in Equation (2), thereby diminishing the overall degradation performance of pollutants [10].

Nanoscale zero-valent iron (nZVI) functions as a substitute for providing Fe²⁺. Within the nZVI/PS environment, Fe²⁺ generation stems from the oxidative degradation of nZVI facilitated by either oxygen-containing or oxygen-depleted conditions, and/or through the influence of PS (as outlined in Equations (3)–(5)) [11]. Diverging from the instantaneous release of Fe²⁺ typical of homogeneous activation methods, nZVI/PS systems enable a more gradual and sustained release of this ion into the water phase. Utilizing this approach minimizes the excessive buildup of Fe²⁺ and its resulting quenching of SO₄^•− (Equation (2)) [12]. Furthermore, Fe²⁺ can be reformed via the recycling of Fe³⁺ at the ZVI’s surface, as illustrated by Equation (6). Therefore, the effectiveness of PS is significantly enhanced [13]. The favorable characteristics of the nZVI/PS procedure have been thoroughly confirmed through prior research.

However, the use of nZVI/PS always presents problems: (i) Iron oxides such as lepidocrocite (γ-FeOOH) and magnetite (Fe₃O₄) form and accumulate on the surface of nZVI, hindering the dissolution and utilization of Fe⁰. (ii) As the degradation reaction proceeds, iron sludge accumulation occurs, and Fe²⁺ is gradually oxidized to Fe³⁺, further inhibiting the generation of SO₄^•−. Thus reducing the efficiency of contaminant degradation [14]. Given the issues with nZVI, it is crucial to study methods to improve its reactivity in applications and fully leverage its advantages in pollutant removal.

The use of organic ligands to enhance the reactivity of nZVI was a significant development in the treatment of organic and inorganic pollutants [15]. L-cysteine (Cys) exists naturally in organisms and has antioxidant, detoxification, and other functions. It has the advantages of a simple preparation process and good water solubility. It is widely used in medicine, food, cosmetics, and other fields [16]. Cys has a large number of groups, such as the thiol, carboxyl, and amino groups; this makes it easy to bind to substrates. For example, Cys can be combined with an iron oxide layer in two ways: one is the coordination with an iron ion after sulfhydryl dissociation, and the other is the deprotonation adsorption of carboxyl [17]. Cys, a type of dissolved organic compound, has demonstrated the ability to suppress the clustering of AgNPs and facilitate the leaching of silver ions from silver-based engineered nanoparticles [18]. Furthermore, it promotes the oxidative dissolution of Fe (III) (hydr)oxides, undergoing transformation into L-cystine (Cy) in the process [19]. Kaden et al. also found that Cys can be used as an electron shuttle to transfer electrons between “Geobacter sulfurreducens” and “Wolinella” succinogenes [17]. Typically, the interaction between pollutants and nZVI takes place at the surface of the latter [20]. As previously mentioned, Cys can alter the characteristics of the nZVI surface, consequently influencing its interaction with contaminants.

In this study, Cys was used to enhance the degradation of MCB in water catalyzed by self-made nZVI, and the reaction mechanism of the Cys/nZVI/PS system was studied. The effects of nZVI dosage, PS concentration, Cys concentration, initial pH value of solution, coexisting ions in water, and other factors on the degradation efficiency of MCB in water by the Cys/nZVI/PS system were studied. The physical and chemical properties and surface morphology of self-made nZVI were analyzed by SEM, XRD, and XPS. The mechanism of degradation of MCB by Cys-enhanced nZVI/PS system and the types of free radicals in the reaction process were speculated by free radical quenching experiment, EPR, XRD, and XPS characterization. This work provides mechanistic insights and operational guidance for Cys-enhanced nZVI-activated PS degradation of chlorobenzene organic pollutants in water.

2. Results and Discussion

2.1. Characterization of nZVI

Chain-like aggregates of spherical particles (Figure 1e,f), exhibiting sizes ranging from 1 to 100 nm (average ≈ 60 nm), were observed via SEM analysis in potassium borohydride-reduced nZVI [21]. Individual particles displayed a spherical morphology (Figure 1c,d) with flake-textured surfaces occasionally showing filamentous impurities (Figure 1a,b).

The particles possess a core–shell structure, featuring metallic iron cores isolated by thin oxide films and encapsulated by an iron oxide shell approximately 2–4 nm thick [22]. X-ray diffraction (XRD) analysis of freshly prepared nZVI showed a weak, broadened peak at 2θ = 44.8° (Figure 1g), corresponding to the (110) plane of body-centered cubic α-Fe⁰. Peak broadening suggests the iron core comprises fine-grained polycrystalline particles [23].

X-ray photoelectron spectroscopy (XPS) of the shell region revealed distinct Fe 2p photoelectron peaks (Figure 1h): Fe 2p_3/2: 711.9 eV (Fe³⁺), 710.4 eV (Fe²⁺), 707.1 eV (Fe⁰); Fe 2p_3/2: 725.5 eV (Fe³⁺), 723.8 eV (Fe²⁺), 719.7 eV (Fe⁰). Peaks at 711.9, 710.4, 725.5, and 723.8 eV confirm the iron oxide shell [24]. The Fe⁰ peak at 707.1 eV verifies the presence of zerovalent iron in the freshly synthesized material. Detection of this Fe⁰ signal indicates an oxide shell thickness below 10 nm. Collectively, these results confirm a core–shell structure consisting of an Fe⁰ core surrounded by an iron oxide layer.

2.2. Reductive Degradation of MCB by Different Systems

A comparison of MCB removal rate in different PS-based Fenton systems at unadjusted pH (5.4–6.7) conditions and 25 °C is shown in Figure 2. Within the 90 min reaction time, both the PS and Cys/PS reaction systems exhibited low removal efficiency for MCB. The simultaneous addition of nZVI/PS, as well as nZVI/Cys/PS, showed a significant removal effect on MCB. At 90 min, the simultaneous addition of nZVI/PS system achieved a removal rate of 77.8%, while the simultaneous addition of nZVI/Cys/PS system achieved a removal rate of 90.1%. However, the addition of only Cys and only nZVI/Cys systems had no effect on the degradation of MCB. These phenomena indicate that the accelerating effect of Cys on MCB degradation is not caused by Cys directly reducing MCB or exchanging electrons with the surface of nZVI alone. We speculate that the presence of Cys increases the opportunity for internal contact between PS and nZVI, thereby increasing the production of free radicals in the system.

The corrosion products of nZVI, such as Fe₃O₄, Fe₂O₂, FeOOH, and Fe₈O₈(OH)₆SO₄, occurred on the solid–liquid interface and acted as a physical barrier preventing the contact of nZVI and PS, which was the main reason why the PS activation was influenced [25]. Two distinct electron transfer pathways operate on passivated nZVI during PS activation: defect-mediated corrosion and semiconductor effects [26]. Corrosion products randomly formed on nZVI surfaces contain structural defects such as grain boundaries or pits [27], enabling direct electron transfer from nZVI to PS. Concurrently, the semiconductor properties of these corrosion products establish electron transfer channels. Fermi energy differences between the metallic iron core and iron oxide or hydroxide shells [28] drive electron migration until Fermi level equilibration [29]. Cys enhances this process through dual mechanisms: the sulfhydryl group (–SH) reduces surface-bound Fe³⁺, increasing defect density and electron donation capacity [30]. Cys adsorbs on nZVI via amino, carboxyl, and sulfhydryl groups, forming Cys-Fe complexes [31,32]. These complexes stabilize nZVI by generating Fe–S bonds, which create delocalized electronic states near the Fermi level [33], thereby bridging Fermi energy disparities. This synergistic effect—accelerating defect generation while establishing electron transfer pathways—explains the significant enhancement of MCB removal in Cys-amended systems.

2.3. The Effect of Different Ligands Enhancing the nZVI/PS System on MCB Degradation

We conducted a systematic investigation on the efficiency of the nZVI/PS system in enhancing the degradation of MCB using five ligands: Cys, citric acid, tartaric acid, oxalic acid, and ethylenediaminetetraacetic acid disodium (EDTA). As shown in Figure 3, at 90 min of reaction, the degradation effect of MCB by the system with different ligands added is: Cys > oxalic acid > citric acid ≈ tartaric acid >> EDTA.

It is worth noting that the enhancement effect of ligands on MCB degradation mainly depends on their chelating and reducing abilities. On the one hand, they form soluble complexes with Fe²⁺/Fe³⁺ to prevent iron precipitation and may alter the reactivity of iron; On the other hand, they can efficiently reduce inert Fe³⁺ to Fe²⁺ with activated PS function, thereby accelerating the Fe²⁺/Fe³⁺ cycle. Cys has both strong reducibility and good chelating properties. Its thiol group (-SH) has extremely strong reducibility and can quickly reduce Fe³⁺ to Fe²⁺, which is its core enhancement mechanism. Simultaneously possessing amino (-NH₂) and carboxyl (-COOH) groups, it can form stable five-membered ring chelates with iron to prevent precipitation of iron hydroxide. The strong reducibility of oxalic acid is highly dependent on UV or visible light irradiation [15]. Our reaction was carried out in a brown reaction bottle in the dark to avoid the photolysis of MCB. This main reduction pathway cannot be carried out, and its effect is greatly reduced, relying solely on complexation to maintain. Citric acid and tartaric acid both have an α-hydroxy acid structure and have a certain reducing ability, which can reduce Fe³⁺ to Fe²⁺, but the reaction rate is much slower than the thiol group of Cys [8]. Their main function is to maintain the solubility and activity of iron in solution through chelation, rather than rapidly driving iron cycling, so the enhancement of the system is mild and slow. EDTA itself does not have strong reducibility. The key is that its complexation constant with Fe³⁺ is much higher than that with Fe²⁺. This will stabilize the Fe³⁺ valence state, but instead hinder the reduction of Fe³⁺ to Fe²⁺, cutting off the critical iron cycle [12]. Meanwhile, EDTA itself competes for free radicals with the target pollutant. So, adding EDTA actually inhibited the degradation of MCB.

In addition to its high reinforcement effect, we emphasize that Cys is a naturally occurring amino acid and a fundamental component of all proteins in living organisms. It can be produced industrially by hydrolyzing human hair or feathers (rich in keratin) [34]. Compared with persistent synthetic chelating agents such as EDTA, not only is the ecological risk lower, but it can also effectively reduce costs in practical applications. Another key point is that Cys is easily biodegradable. This greatly reduces the likelihood of its persistence and accumulation in downstream environments, minimizing long-term exposure risks. The product of Cys oxidation, Cy ester, is also a natural amino acid dimer with low ecological toxicity and can also be biodegraded [19].

2.4. The Influence of Different Reaction Conditions on MCB Degradation

A number of degradation experiments were conducted to investigate the impact of different reaction conditions on the degradation of MCB. Figure 4 illustrates the degradation of MCB under varying doses of nZVI (Figure 4a), PS (Figure 4b), and Cys (Figure 4c), as well as different initial pH levels (Figure 4d).

The nZVI acts as a vital alternative iron source for providing divalent iron during the activation of persulfate to produce active species that decompose MCB. Consequently, the quantity of nZVI incorporated directly determines the system’s efficacy in eliminating MCB. Consequently, four dosage levels of nZVI were employed in this research: 0.05 g/L, 0.075 g/L, 0.1 g/L, and 0.15 g/L. As illustrated in Figure 4a, following 90 min of reaction, the chlorobenzene degradation efficiencies at these concentrations were observed to be 56.5%, 68.6%, 90.7%, and 85.7%, respectively. As the concentration of nZVI rose from 0.05 g/L to 0.075 g/L and subsequently to 0.1 g/L, the removal efficiency of MCB exhibited a corresponding increase. Nevertheless, when the nZVI dosage reached 0.15 g/L, the MCB degradation rate was observed to decline relative to the 0.1 g/L concentration. A surplus of nZVI directly interacts with SO₄^•− within the system, leading to free radical scavenging (Equation (2)) and diminishing the efficiency of MCB degradation. Furthermore, overabundant nZVI prompts the formation of a thicker passivation layer, where Fe³⁺/Fe²⁺ swiftly precipitate as an oxidized coating over the nZVI surface, impeding electron transfer [35].

Further investigations assessed how varying the PS concentration influenced the efficiency of the removal of MCB. In instances where the PS concentration was absent, utilizing only nZVI resulted in no detected loss of MCB within the solution, suggesting that minimal adsorption and/or reduction processes involving MCB took place on the nZVI surface (refer to Figure 3). Increasing the initial PS concentration favored MCB oxidation in the nZVI/PS/Cys process as well (Figure 4b). For instance, at the initial PS concentration of 1 mM, approximately 65.4% MCB was eliminated in 90 min. When the initial PS increased to 3 mM, the removal increased to approximately 91.2% over the same time period (Figure 4b). Previous studies have demonstrated that the limiting step in SO₄^•− generation was the surface reaction between nZVI and the oxidant [36]. Thus, the higher PS concentration increased the corrosion of nZVI (Equation (5)) and generated more Fe²⁺, which, in combination with the higher PS concentration per se, resulted in faster generation of SO₄^•− (Equation (1)) [37]. Thus, degradation of MCB was accelerated.

The overall MCB decomposition could be attributed to the reactions with all oxidizing species (e.g., SO₄^•− and HO^•) present in the system. SO₄^•− is generally the predominant radical species responsible for the degradation of organics in activated PS oxidation systems at pH < 9 [38,39]. However, other oxidizing species, such as HO^•, could be produced as well (Equations (9) and (10)) [40] and contributed to the oxidation of MCB. When the concentration of PS increased to 4 mM, it was found that the improvement in the degradation rate of MCB was almost negligible compared to a PS concentration of 3 mM. This may be due to the inhibition of SO₄^•− generation as nZVI is depleted in the system (Equations (1) and (5)), and excessive SO₄^•− may dimerize through a self-quenching reaction to generate persulfate ions (Equation (11)).

The corrosion products coated on the nZVI surface prevented the contact of the inner nZVI with PS to generate more radicals. Therefore, the presence of Cys increases the chances of PS contacting Fe⁰ in nZVI, thereby increasing the yield of free radicals in the system [41]. It can be clearly seen from Figure 4c that the system with 0.2 mMCys added but without nZVI has an extremely poor degradation effect on MCB, which once again proves that Cys cannot activate PS and has no degradation effect on MCB itself. Instead, it reduces the iron oxide on the surface of nZVI to provide more active sites for PS. Additionally, within a certain range, as the concentration of Cys increases, the degradation effect of MCB also improves. When the concentration of Cys is 0.1 mM, the degradation efficiency of MCB reaches 92.6%. When the concentration of Cys was high, it competes with MCB to consume the SO₄^•− produced by the nZVI/PS system [42]. When the concentration of Cys is 0.2 mM, its efficiency in degrading MCB is not as good as when the concentration of Cys is 0.05 mM.

The effect of initial pH on MCB removal efficiency is shown in Figure 4d. At different initial pH values, the nZVI/PS/Cys system exhibits significant differences in MCB removal efficiency, with an overall trend of decreasing MCB removal rate with increasing initial pH. When the pH is 3 and 5, the degradation rates of MCB are 88.2% and 82.6%, respectively. When the pH is low, a large amount of H⁺ in the solution can accelerate the corrosion of Fe⁰, activate surface activity, and release more charged particles to participate in the reaction [43]. On the other hand, under acidic conditions, the deposition of iron oxide on the nZVI surface is inhibited, causing the adsorption sites to be continuously exposed to the reaction system, as shown in chemical reaction Equation (2). When the pH is 7 and 9, the degradation rates of MCB are 78.7% and 64.7%, respectively. Higher pH values will inhibit the reaction between nZVI and H⁺, as well as the surface hydrolysis to produce ferrous ions (Equations (3) and (4)), and also lead to the formation of hydroxide precipitation of ferrous ions or iron ions (Equations (12) and (13)). The reduction of ferrous ions will lead to a decrease in the activation of persulfate ions, thereby inhibiting the production of SO₄^•−.

2.5. The Influence of Coexisting Ions in Water on the Degradation of MCB

In the water treatment research of advanced oxidation technology, chloride ion (Cl⁻) and bicarbonate ion (HCO₃⁻) are often selected as the research objects of the influence of coexisting ions, mainly based on their universality, high concentration, strong reactivity in natural water and industrial wastewater, and their key interference mechanism on the free radical path.

In order to investigate its effect on the degradation of MCB in water by Cys-enhanced nZVI activated PS system, this study conducted five comparative experiments with Cl⁻ concentrations of 0 mM, 5 mM, 10 mM, 50 mM, and 100 mM. The specific experimental results are shown in Figure 5a. Overall, the effect of chloride ions on the degradation of MCB in the entire system is relatively small, but there is also a slight impact. After 90 min of reaction, the degradation rates of MCB in five systems with Cl⁻ concentrations of 0 mM, 5 mM, 10 mM, 50 mM, and 100 mM reached 90.2%, 90.1%, 88.8%, 83.2%, and 79.3%, respectively. It can be seen that compared to the system without Cl⁻, a small amount of Cl⁻ enhances the degradation ability of MCB within 30 min. However, as the reaction time increases, the degradation efficiency of MCB at 90 min is not significantly affected. Cheolyong Kim et al. found that the oxidation rate of phenol significantly increases in the early stages of the reaction in the presence of 25–200 mM Cl⁻ [44]. They believe that this is because excessive sulfate ions (SO₄^•−) react with Cl⁻ to produce short-lived reactive chlorine species, such as ^•Cl and ^•Cl₂⁻, rather than being cleared by Fe²⁺ or other SO₄^•− [44]. When the concentration of Cl⁻ continued to increase to 50 mM and 100 mM, the degradation rates of MCB in these two systems decreased by 7% and 10.9%, respectively, compared to the system without Cl⁻ addition. Similarly, in the study by Cheolyong Kim et al., the oxidation rate of phenol decreased at high Cl⁻ concentrations (>400 mM) [44]. This may be because Cl⁻ reacts with the strong oxidant SO₄^•−, causing SO₄^•− to lose electrons and generate SO₄²⁻, which has no contribution to the degradation of MCB. Cl⁻ will be oxidized to ^•Cl, and the newly generated ^•Cl will continue to react with Cl⁻ to generate ^•Cl₂⁻, which is unstable and will react with Cl₂ and Cl⁻, thereby reducing the ability of the entire system to degrade MCB. The oxidation capacity of ^•Cl is weaker than that of SO₄^•−. The specific reaction process is shown in Equations (14)–(16).

Bicarbonate ion (HCO₃⁻) is an inevitable component in water, which can significantly inhibit the ability of the system to degrade MCB. Bicarbonate ion can be decomposed into carbonate ion. The acidity and alkalinity of natural water are all regulated by the HCO₃⁻/CO₃²⁻ ion pair. In this study, four comparative experiments were conducted with HCO₃⁻ concentrations of 0 mM, 1 mM, 2 mM, and 5 mM. The specific experimental results are shown in Figure 5b. After 90 min of reaction, the degradation rates of MCB by the four systems with HCO₃− concentrations of 0 mM, 1 mM, 2 mM, and 5 mM were 88.2%, 76.3%, 56.6%, and 39.5%, respectively. As the concentration of HCO₃⁻ increased, the degradation rates of MCB in the 1 mM, 2 mM, and 5 mM systems decreased by 11.9%, 31.6%, and 48.7%, respectively. There may be two reasons for this. Firstly, the addition of HCO₃⁻ leads to an increase in the pH value of the entire system. As mentioned earlier, pH has an important impact on the entire system, and with the increase in pH, the degradation rate of MCB by the system will also decrease. Secondly, with the addition of HCO₃−, SO₄^•− in the solution will react with it to generate sulfate and bicarbonate radicals, which can then be further converted into hydrogen ions and less oxidizing bicarbonate radicals [45]. The specific reaction is shown in Equations (17) and (18).

2.6. Efficiency of nZVI/Cys/PS System in Degrading MCB in Different Actual Water Environments

As shown in Figure 6, although the initial pH was uniformly optimized to 3, the degradation efficiency of different water bodies showed significant differences, in the order of: ultrapure water (UW) ≈ groundwater (GW) > tap water (TW) > surface water (SW) > Secondary biochemical effluent of coking wastewater (SBECW).

The two degradation curves of GW and UW are very close. GW promotes corrosion due to the presence of trace amounts of Cl⁻, and the initial degradation efficiency is even slightly faster than UW, but the final efficiency is comparable after 90 min. Due to the passivation effect of residual chlorine and phosphate on Cys and nZVI in TW, the initial degradation efficiency is directly reduced, but the system still maintains a certain degree of degradation ability. The high concentration of NOM in SW continuously competes with MCB for active species throughout the entire reaction process, resulting in a slow and low degradation efficiency [4]. However, the SBECW of coking wastewater contains high salt content, and the synergistic effect of various complex organic matter causes free radicals to be consumed as soon as they are generated [5]. The nZVI surface is severely passivated, making it difficult for the reaction to proceed effectively. This trend clearly indicates that water matrix components such as natural organic matter (NOM), phosphates, ammonia nitrogen, etc., have a strong inhibitory effect on the Cys/nZVI/PS system. The degradation curve in GW highly overlaps with UW, demonstrating the enormous potential of this technology in GW resource restoration [8]. However, for SW containing high concentrations of NOM and SBECW with complex components, the system performance sharply decreases, indicating the need for appropriate pretreatment to remove interfering substances before practical application.

2.7. Analysis of Active Substances in the Degradation of MCB in Water by nZVI/Cys/PS System

Advanced oxidation processes based on persulfate, tertiary butanol (TBA), and methanol (MeOH) are commonly used as radical scavengers. They react rapidly with radicals, and their reaction products are inert, generally not interfering with the chain reaction process of the Fenton process [46]. Therefore, many researchers use TBA as a scavenger for hydroxyl radicals when analyzing active substances in advanced oxidation technologies, while MeOH is a co-scavenger for both hydroxyl radicals and sulfate radicals. After adding free radical scavengers to the reaction system, the dominant active substances in the system can be analyzed by changes in the removal efficiency of organic pollutants.

As shown in Figure 7a, as analyzed above, TBA is a scavenger for hydroxyl radicals, while MeOH is a co-scavenger for both hydroxyl radicals and sulfate radicals. Therefore, the capture effect after adding MeOH is significantly better than TBA. Within 90 min, the degradation rates of MCB were 74.6%, 50.3%, 70.5%, 24.3%, and 89.4% in five control experiments with and without the addition of 100 mM TBA, 100 mM MeOH, 200 mM TBA, and 200 mM MeOH, respectively. Compared with the control group without any capture agent, the addition of 100 mM TBA, 100 mM MeOH, 200 mM TBA, and 200 mM MeOH decreased by 14.6%, 39.1%, 18.9%, and 65.1%, respectively. The degradation rate of MCB by introducing 100 mM TBA and MeOH differed by 24.3%, while increasing the concentration to 200 mM resulted in a difference of 46.2%. It can be seen that although HO^• also plays a certain role in the degradation of MCB, it is obvious that both free radicals can capture MeOH, but the quenching effect of TBA capturing only HO^• is not significant. Therefore, SO₄^•− plays a dominant role in the degradation of MCB in water by the Cys-enhanced nZVI/PS system.

In order to identify the existence of radicals generated, the EPR experimental analysis was carried out by using DMPO as the spin-trapping agent. The EPR results indicated that, as shown in Figure 7b, DMPO-SO₄ (six lines, 1:1:1:1:1:1, blue heart-shaped) and DMPO-OH (four lines, 1:2:2:1, red rhombus) signals were captured, which proved the existence of SO4^•− radicals and HO^• radicals in nZVI/Cys/PS system [47,48].

2.8. Proposed Intermediates and Pathways for MCB Degradation

While the primary focus of the present study was on elucidating the role of Cys in enhancing the radical generation and iron cycling, we did not perform new GC-MS analysis for intermediate identification in this specific setup. To assess the potential formation of toxic intermediates and to gain a deeper understanding of the degradation mechanism, based on the identified reactive oxygen species (SO₄^•− and HO^•) and our team’s detailed analysis of the pathway and intermediate products of sulfate radical oxidation degradation of MCB in previous studies, a proposed degradation pathway and intermediate products for MCB in the Cys/nZVI/PS system were proposed [1,2].

The degradation is anticipated to initiate through two primary routes: (i) the preferential attack of sulfate radicals (SO₄^•−) on the aromatic ring, leading to the formation of hydroxylated intermediates, and (ii) the direct oxidative cleavage of the C-Cl bond. The hydroxylation of MCB results in the formation of various chlorophenol isomers (e.g., 2-chlorophenol and 4-chlorophenol), which were consistently identified as the first key intermediates in our previous studies on Fe²⁺/Cys/PS systems [1,2].

Subsequently, these chlorophenols undergo further oxidation. This stage involves repeated hydroxylation, leading to the formation of chlorinated dihydroxybenzenes (e.g., chlorocatechol), and concomitant dechlorination reactions, yielding non-chlorinated polyhydroxylated benzenes such as hydroquinone and catechol. These compounds are highly susceptible to oxidation and can be readily converted to benzoquinone.

The cleavage of the aromatic ring of these phenolic intermediates marks the crucial step towards complete mineralization. The ring opening, likely initiated by the attack of radicals on the C-C bond adjacent to the hydroxyl group, generates a series of small, straight-chain carboxylic acids. Common products identified in this stage include maleic acid, fumaric acid, oxalic acid, and formic acid [1]. Finally, these low-molecular-weight organic acids are mineralized to CO₂ and H₂O.

A schematic illustration of this proposed pathway is presented in Figure 8. It is important to note that under this pathway, the degradation process of MCB is also a detoxification process, as all intermediates have reduced acute and chronic toxicity to fish, water fleas, and green algae compared to the parent compound. In addition, the potential for bioaccumulation has also decreased, and the bioaccumulation coefficients of all intermediates are relatively low [2]. While the formation of toxic intermediates (e.g., chlorophenols, quinones) is possible, the presented pathway demonstrates their transitory nature. The powerful oxidizing environment of the Cys/nZVI/PS system ensures the rapid further oxidation and ultimate destruction of these compounds, thereby mitigating the risks of secondary toxicity and leading to efficient mineralization of the parent pollutant.

2.9. Mechanism of Cys Enhanced the Degradation of MCB by nZVI/PS System

To verify the claim in Section 2.2 that Cys enhances the surface electron transfer of nZVI through a dual mechanism, thereby activating the degradation of MCB by PS, we characterized nZVI after different system reactions using XRD and XPS. As shown in Figure 9a, an XRD diffractogram of the reacting nZVI in the absence of Cys revealed the appearance of peaks of magnetite (Fe₃O₄) at 37.1°, 53.4°, and 62.5° 2θ, and those of α-Fe⁰ at 44.7° and 65.0° 2θ. These results indicate that in the absence of Cys, the main component of the iron oxide layer formed on the surface of nZVI after the reaction is Fe₃O₄. However, in the presence of Cys, the XRD analysis of the reacting nZVI showed an increase in peaks of lepidocrocite (γ-FeOOH, 26.4°, 41.2°, 60.2°, and 67.9° 2θ), magnetite (Fe₃O₄, 37.1°, 53.4°, and 62.5° 2θ), and α-Fe0 (44.7° and 65.0° 2θ). Furthermore, the peak of α-Fe⁰ decreased significantly in intensity compared with the nZVI that had not reacted (Figure 1g). In previous research, Yoon et al. found that in the presence of Cys, the morphology of nZVI undergoes significant changes due to the appearance of lepidocrocite in the iron oxide layer, presenting as sheet-like particles composed of needle-like structures, effectively improving the agglomeration phenomenon of nZVI [49]. The TEM characterization of nZVI after different system reactions can refer to Figure A1.

The characterization results of XPS also support this claim; the spectrum of Fe (2p_3/2) in the narrow region has four peaks attributed to the binding energies of Fe⁰ (707.1 eV), Fe²⁺ (710.4 eV), and Fe³⁺ (711.9 eV), respectively. As shown in Figure 9b, in the absence of Cys, the total iron content in the remaining nZVI after the reaction is 58.6% of the initial nZVI. The content of Fe³⁺ (34.9%) is similar to that of the initial nZVI (33.9%), but the content of Fe⁰ decreases from 15.8% to 7.3%, and the content of Fe²⁺ increases from 50.3% to 57.8%. In the presence of Cys, XPS analysis showed that the total iron content in the remaining nZVI after the reaction was only 29.88% of the initial nZVI, with a significant increase in Fe²⁺ (63.2%) and a sharp decrease in Fe⁰ (0.3%) and Fe³⁺ (36.5%). These results support our hypothesis that Cys enhances electron transfer between nZVI and MCB, thereby accelerating the redox reaction between the Fe⁰ nucleus and PS. Furthermore, we can confirm that the dissolution of the oxide layer and the exposure of more active sites are not the main reasons for Cys enhancing the reactivity of nZVI in the system [50]. The above characterization analysis supports our hypothesis that the reduction degradation of MCB is promoted by the Cys-mediated electron transfer activation of PS on the nZVI surface.

In addition, Cys can also accelerate the redox cycle of Fe³⁺/Fe²⁺ in the nZVI/PS system. Doong R A et al. demonstrated that cysteine can serve as an electron carrier and transfer electrons to Fe³⁺ [51]. Chengdu Qi et al. found that cysteine, as a mild reducing agent, can reduce dissolved Fe³⁺ to Fe²⁺ [52]. Li et al. found that adding cysteine to the classical Fenton system can greatly enhance the degradation of pollutants, demonstrating that cysteine can effectively reduce Fe³⁺ to Fe²⁺ [53]. The conversion between Cys and Fe²⁺ opens up a new shortcut channel, and the possible reactions that may occur after the introduction of Cys are shown in Equations (19)–(22).

Based on the above results and analysis, Figure 10 demonstrates a potential reaction mechanism. We believe that the reduction degradation of MCB is promoted by Cys-mediated electron transfer activation of PS. Through pathway I, Cys present in the system transfers electrons to Fe³⁺ for reduction, which is then oxidized to form L-cystine (Cy). Fe²⁺ is then oxidized to Fe³⁺ by PS, forming a Fe³⁺/Fe²⁺ cycle. Using nZVI as the electron donor, the obtained Cy is reduced to Cys to form a Cys/Cy cycle. Cys adsorbs onto nZVI through amino (-N₂H), carboxyl (-COOH), and thiol groups (-SH), forming a Cys Fe complex that further releases Fe²⁺ (pathway II). Through pathway III, the -SH in Cys reduces the trivalent iron in the surface iron oxide of nZVI, increases defect density and electron donor capacity, allowing electrons to transfer directly from nZVI to PS and release Fe²⁺. As mentioned above, Cys acts as an electric shuttle in the reaction, providing sufficient Fe²⁺ to PS through various pathways to promote the production of SO₄^•− and significantly improve the degradation efficiency of MCB.

Another thing worth noting is that the presence of Fe⁰ and Fe²⁺ species even after the reaction indicates that the core–shell structure is maintained and the catalyst is not fully passivated. More importantly, the Cys layer effectively mitigates the formation of a thick, inert Fe (III) oxide shell, which is the primary cause of nZVI deactivation. This suggests that the Cys/nZVI/PS system has inherent potential for longer-term activity. Future studies can focus on optimizing reactor configurations for continuous operation and catalyst recycling.

2.10. Formatting of Mathematical Components

The equations mentioned in this article are as follows:

$${\mathrm{S}}_2{\mathrm{O}}_8^{2-}+{\mathrm{Fe}}^{2+}\to {\mathrm{Fe}}^{3+}+{\mathrm{SO}}_4^{•-}+{\mathrm{SO}}_4^{2-}$$

(1)

$${\mathrm{SO}}_4^{•-}+{\mathrm{Fe}}^{2+}\to {\mathrm{Fe}}^{3+}+{\mathrm{SO}}_4^{2-}$$

(2)

$$2{\mathrm{Fe}}^0+{\mathrm{O}}_2+2{\mathrm{H}}_2\mathrm{O}\to 2{\mathrm{Fe}}^{2+}+4\mathrm{O}{\mathrm{H}}^{-}$$

(3)

$${\mathrm{Fe}}^0+2{\mathrm{H}}_2\mathrm{O}\to {\mathrm{Fe}}^{2+}+2{\mathrm{OH}}^{-}+{\mathrm{H}}_2$$

(4)

$${\mathrm{S}}_2{\mathrm{O}}_8^{2-}+{\mathrm{Fe}}^0\to {\mathrm{Fe}}^{2+}+2{\mathrm{SO}}_4^{2-}$$

(5)

$${\mathrm{Fe}}^0+2{\mathrm{Fe}}^{3+}\to 3{\mathrm{Fe}}^{2+}$$

(6)

$$4{\mathrm{Fe}}^{3+}+3{\mathrm{BH}}_4^{-}+9{\mathrm{H}}_2\mathrm{O} \to 4{\mathrm{Fe}}^0\downarrow +3{\mathrm{H}}_2{\mathrm{BO}}_3^{-}+12{\mathrm{H}}^{+}+6{\mathrm{H}}_2\uparrow$$

(7)

$${\mathrm{Fe}}^0+2{\mathrm{H}}^{+}\to {\mathrm{Fe}}^{2+}+{\mathrm{H}}_2$$

(8)

$${\mathrm{SO}}_4^{•-}+{\mathrm{HO}}^{-}\to {\mathrm{HO}}^{•}+{ \mathrm{SO}}_4^{2-}$$

(9)

$${\mathrm{SO}}_4^{•-}+{\mathrm{H}}_2\mathrm{O}\to {\mathrm{HO}}^{•}+{\mathrm{SO}}_4^{2-}+{\mathrm{H}}^{+}$$

(10)

$$2{\mathrm{SO}}_4^{•-}\to {\mathrm{S}}_2{\mathrm{O}}_8^{2-}$$

(11)

$${\mathrm{Fe}}^{2+}+2{\mathrm{OH}}^{-}\to {\mathrm{Fe}\left(\mathrm{OH}\right)}_2$$

(12)

$${\mathrm{Fe}}^{3+}+3{\mathrm{OH}}^{-}\to {\mathrm{Fe}\left(\mathrm{OH}\right)}_3$$

(13)

$${\mathrm{SO}}_4^{•-}+{\mathrm{Cl}}^{-}\to {}_{ }{}^{•}\mathrm{Cl}+{\mathrm{SO}}_4^{2-}$$

(14)

$${}_{ }{}^{•}\mathrm{Cl}+{\mathrm{Cl}}^{-}\to {}_{ }{}^{•}{\mathrm{Cl}}_2^{-}$$

(15)

$${}_{ }{}^{•}{\mathrm{Cl}}_2^{-}+{}_{ }{}^{•}{\mathrm{Cl}}_2^{-}\to 2{\mathrm{Cl}}^{-}+{\mathrm{Cl}}_2$$

(16)

$${\mathrm{SO}}_4^{•-}+{\mathrm{HCO}}_3^{-}\to {\mathrm{SO}}_4^{2-}+{\mathrm{HCO}}_3^{•}$$

(17)

$${\mathrm{HCO}}_3^{•}\to {\mathrm{H}}^{+}+{\mathrm{CO}}_3^{•-}$$

(18)

$$\mathrm{CSH} \leftrightarrow {\mathrm{CS}}^{-}+{\mathrm{H}}^{+}$$

(19)

$${\mathrm{Fe}}^{3+}+{\mathrm{CS}}^{-}\to {\left[\mathrm{Fe}\right(\mathrm{CS}\left)\right]}^{2+}$$

(20)

$${\left[\mathrm{Fe}\right(\mathrm{CS}\left)\right]}^{2+}\to {\mathrm{Fe}}^{2+}+{\mathrm{CS}}^{•}$$

(21)

$$2{\mathrm{CS}}^{•}\to \mathrm{CSSC}$$

(22)

3. Materials and Methods

3.1. Materials

Iron trichloride hexahydrate (FeCl₃·6H₂O, 99%), ethanol absolute (C₂H₆O, 99.7%), potassium borohydride (KBH₄, 94%), L-cysteine (C₃H₇NO₂S, 98.5%), mono-chlorobenzene (MCB, C₆H₅Cl, 99%), citric acid (C₆H₈O₇, 99.5%), ethylenediaminetetraacetic acid disodium (C₁₀H₁₄N₂Na₂O₈, 99%) were obtained from Sinopharm Chemical Reagent Co. Ltd., Shanghai, China. Sodium persulfate (PS, Na₂S₂O₈, 94%), tartaric acid (C₄H₆O₆, 99%), and oxalic acid (C₂H₂O₄·2H₂O, 99.5%) were purchased from Shanghai Macklin Biochemical Co., Ltd., Shanghai, China.

GW was taken from a drinking water well in Jiyuan City, Henan Province. TW was collected from the laboratory water supply system without further treatment. SW was taken from the lake water on the campus of Henan University of Technology. SBECW was taken from a coking plant sewage treatment station in Pingdingshan City, Henan Province.

3.2. Synthesis of Nano Zero-Valent Iron

Nano zero-valent iron (nZVI) was synthesized via a liquid-phase reduction method (see Figure 11 for the experimental setup). The procedure was as follows: FeCl₃·6H₂O (9.6534 g) was precisely weighed using an analytical balance (0.0001 g precision) and placed into a three-necked round-bottom flask. Anhydrous ethanol (40 mL) and ultrapure water (10 mL) were sequentially added to the flask, and the mixture was stirred uniformly using a magnetic stirrer. Separately, KBH₄ powder (1.0788 g) was dissolved in ultrapure water (100 mL) to prepare a 0.2 mol/L KBH₄ solution, which was used immediately after preparation.

A continuous N₂ flow was introduced into one neck of the flask. The freshly prepared KBH₄ solution was then slowly added to the flask using a peristaltic pump under vigorous stirring. After complete addition, stirring was continued for 30 min, resulting in the formation of a significant amount of black solid. The reaction mixture was then washed three times with 50 mL of 30% ethanol (v/v), followed by three washes with 50 mL of absolute ethanol. The purified solid product was stored overnight in a freezer at −20 °C and subsequently lyophilized in a freeze-dryer. The reaction is represented by Equation (7).

3.3. Degradation Experiments MCB

A 3 mM MCB stock solution was prepared by injecting 150 μL of pure MCB into a 500 mL amber volumetric flask with ultrapure water dilution. The solution was sealed with parafilm and magnetically stirred overnight for complete dissolution and then transferred to a 100 mL amber syringe wrapped in aluminum foil. The syringe needle port was secured with a rubber septum and sealing tape. All MCB handling used fine-gauge needles to minimize volatilization.

For reactions, pre-weighed fresh nZVI was added to 24 mL amber headspace vials (PTFE/silicone septum). For enhanced systems, a defined volume of Cys solution was introduced. Subsequently, MCB and persulfate solutions were sequentially injected. The total volume was adjusted to 24 mL with ultrapure water. The final MCB concentration in the solution was 0.3 mM.

Sealed vials were vortex-mixed at timed intervals. At designated timepoints, 10 mL aliquots were collected via needle-free syringe and immediately quenched in amber vials containing 3 M sodium thiosulfate/NaCl solution for MCB analysis.

We conducted additional experiments in GW, TW, SW, and STW to evaluate the practical applicability of our system. All water samples were filtered through 0.45-μm membranes prior to use to remove suspended solids, and the initial pH was adjusted to 3 using diluted H₂SO₄ for experimental consistency. In addition, the effects of different ligands such as citric acid, oxalic acid, tartaric acid, and disodium ethylenediaminetetraacetic acid on the degradation of chlorobenzene in the experimental system were compared to determine the enhancing effect of Cys.

3.4. Characterization

By observing the surface morphology and microstructure of nZVI using scanning electron microscopy (SEM, Hitachi SU8010, Hitachi High-Tech Science Corporation, Tokyo, Japan), crystal structure information of nZVI (fresh), nZVI (used), and nZVI/Cys (used) was acquired by powder X-ray diffractometer (XRD, Smart Lab, 9 kW, Hitachi High-Tech Science Corporation, Tokyo, Japan) with Cu Kα radiation. Characterization of elemental compositions and chemical states of nZVI surfaces by X-ray photoelectron spectroscopy (XPS) analysis (ESCALAB 250Xi, Thermo Fisher Scientific, East Grinstead, United Kingdom). The available reactive oxygen species (ROS) generated in the nZVI/PS/Cys process were detected by electron paramagnetic resonance (EPR) spectroscopy (X-band BRUKER EMX, Bruker Corporation, Billerica, MA, USA) using 5,5-dimethyl-1-pyrroline N-oxide (DMPO) as a spin-trapping agent.

3.5. Analytical Methods

The concentration of MCB was analyzed using a flame ionization detector (FID) and an HP-5 column (30 m × 0.32 mm × 0.25 μm) on a gas chromatograph (GC, HP 6890 Series, Hewlett Packard Co. Ltd., Palo Alto, CA, USA). The specific steps were as follows: Add a certain volume of n-hexane to the headspace bottle containing the test solution for extraction, shake vigorously for 15 min, let it stand, and use a 1 mL injection needle to extract 1 mL of the supernatant into a 2 mL short threaded sample bottle for machine testing. The detection limit was 0.006 mg/L. To ensure data quality, all experiments were performed in triplicate, and the average values were used to fit curves.

4. Conclusions

This study establishes Cys as a highly effective enhancer for nZVI/PS systems in degrading MCB in water. Key findings are summarized as follows:

1. Acidic conditions (pH = 3) maximized MCB degradation efficiency (88.2% in 90 min), leveraging H⁺-accelerated nZVI corrosion for sustained Fe²⁺ release and SO₄^•− generation. Alkaline conditions (pH = 9) reduced efficiency to 64.7% due to Fe (OH)_x precipitation inhibiting iron cycling. Optimal performance occurred at nZVI: 0.1 g/L, PS: 3 mM, and Cys: 0.1 mM, beyond which aggregation (nZVI) or radical scavenging (Cys) reduced efficacy.

2. Low concentrations of Cl⁻ (≤10 mM) slightly enhanced degradation via pitting corrosion, while >50 mM suppressed efficiency (79.3% at 100 mM and 90.2% control) by generating less-reactive chlorine radicals (^•Cl/^•Cl₂⁻). HCO₃⁻ strongly inhibited degradation even at 1 mM (76.3% efficiency), with severe suppression at 5 mM (39.5%), primarily due to SO₄^•− scavenging, yielding weak oxidants (HCO₃^•/CO₃^•−).

3. MCB degradation is driven by both SO₄^•− and HO^• simultaneously. Cys enhances the reduction degradation of MCB by acting as an electron shuttle in the nZVI/PS system. It transfers electrons to Fe³⁺ for reduction, forming a Fe³⁺/Fe²⁺ cycle. Using nZVI as an electron donor, a Cys Fe complex is formed to further release Fe²⁺. By reducing the trivalent iron in the surface oxide of nZVI through -SH, the defect density and electron donor capacity increase, allowing electrons to transfer directly from nZVI to PS and release Fe²⁺.

The Cys/nZVI/PS system offers a sustainable strategy for chlorinated pollutant remediation, leveraging Cys to overcome intrinsic limitations of iron-based AOPs. Critical operational thresholds (pH, ion concentrations) identified herein provide essential guidance for real-water applications. Additionally, it is worth noting that high concentrations of PS released pose a threat to aquatic organisms, and the production of sludge also causes secondary pollution to a certain extent. Therefore, in practical applications, if necessary, established post-treatment processes such as gypsum precipitation (CaSO₄) or sludge removal devices in subsequent treatment steps can be used to effectively control the level of effluent sulfate and sludge formation, and improve the practical applicability of the Cys/nZVI/PS system for degrading MCB in water.