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Article

Activation of ClO2 by Nanoscale Zero-Valent Iron for Efficient Soil Polycyclic Aromatic Hydrocarbon Degradation: New Insight into the Relative Contribution of Fe(IV) and Hydroxyl Radicals

by
Xiaojun Hu
1,
Xiaorong Xing
1,
Fan Zhang
1,
Bingzhi Li
1,
Senlin Chen
1,
Bo Wang
1,
Jiaolong Qin
1,* and
Jie Miao
2,*
1
School of Chemical and Environmental Engineering, Shanghai Institute of Technology, Shanghai 201418, China
2
School of Environmental Science and Engineering, Nanjing Tech University, Nanjing 211816, China
*
Authors to whom correspondence should be addressed.
Toxics 2025, 13(1), 36; https://doi.org/10.3390/toxics13010036
Submission received: 23 November 2024 / Revised: 22 December 2024 / Accepted: 30 December 2024 / Published: 5 January 2025
(This article belongs to the Special Issue Novel Remediation Strategies for Soil Pollution)
Browse Figures
Versions Notes

Abstract

:
Recently, the activation of chlorine dioxide (ClO2) by metal(oxide) for soil remediation has gained notable attention. However, the related activation mechanisms are still not clear. Herein, the variation of iron species and ClO2, the generated reactive oxygen species, and the toxicity of the degradation intermediates were explored and evaluated with nanoscale zero-valent iron (nFe0) being employed to activate ClO2 for soil polycyclic aromatic hydrocarbon (PAH) removal. With an optimized ClO2/nFe0 molar ratio of 15:1 and a soil/water ratio of 3:1, the degradation efficiency of phenanthrene improved 12% in comparison with that of a ClO2-alone system. The presence of nFe0 significantly promoted ClO2 consumption (improved 85.4%) but restrained ClO2− generation (reduced 22.5%). The surface Fe(II) and soluble Fe(II) in the ClO2/nFe0 system was 2.0-fold and 2.8-fold that in the nFe0 system after 2 min. Electron paramagnetic resonance analysis, along with quenching experiments, revealed that Fe(IV), HOCl, and •OH dominated phenanthrene degradation in a ClO2/nFe0 system, with oxidation contributions, respectively, of 34.3%, 52.8% and 12.9%. The degradation intermediates of PAHs in the ClO2/nFe0 system had lower estimated toxicity than those of the ClO2 system. The lettuces grown in ClO2/nFe0-treated soil displayed better results in bioassay indexes than those grown in ClO2-treated soil. This study offers new perspectives for the remediation of organic-pollutant-contaminated soil by using metal-activated ClO2 technology.

1. Introduction

Polycyclic aromatic hydrocarbons (PAHs), many of which are human carcinogenic, teratogenic, and ecotoxic, are generated when organic matter and fossil fuels incompletely combust [1]. PAHs pose serious threats to human health and ecological systems by way of amplification effects in the food chain [2,3]. Sixteen parent PAHs have been grouped into the priority pollutants category by the U.S. Environmental Protection Agency [4]. Due to their hydrophobic nature and persistence, PAHs can be found ubiquitously in the environment, with soil being their ultimate sink [5,6]. The removal of PAHs from soil is a challenge, because PAHs are strongly adsorbed into soil organic matter and are encapsulated in soil minerals [6,7]. Thus, developing effective methods to remediate PAH-contaminated soil is urgently necessary.
As a one-electron oxidant (E0 = 0.936 V), chlorine dioxide (ClO2) has the advantage of having less pH dependence and lower dissolved organic form disinfection byproduct formation potential in comparison with free chlorine. ClO2-based advanced oxidant processes have been widely applied in water and wastewater treatment [8,9,10,11]. Peng et al. [8] reported reactive species (ClO•, Cl•, •OH, and ozone) generated when ClO2 was activated by UV radiation in the UVA range. Su et al. [11] reported that active ClO2 radicals and photo-induced •OH radicals were produced in a combined ClO2-photocatalysis system, which significantly increased the degradation rate of norfloxacin. Gaseous ClO2 has been proven to have an antimicrobial effect on pathogenic bacteria and has been used as a substitute soil fumigant instead of methyl bromide [12,13]. Our previous work showed that ClO2 was an effective oxidizing agent for PAHs in soil [14]. The mechanism involves ClO2 attacking the atoms of PAHs which have the strongest electron-donating ability. However, developing ClO2-based technologies for soil remediation is restricted by lengthy reaction times, low treatment efficiency, high selectivity for organic pollutants, and high costs with huge ClO2 dosages [15,16]. Recently, the activation of ClO2 by metal or metal oxide has gained a large amount of interest. Shi et al. [17] employed carbon–MnO2 to catalyze ClO2 for the enhanced removal of o-chlorophenol from wastewater. Ma et al. [15] reported that montmorillonite-supported Fe3O4-CuO exhibited excellent catalytic activity in the ClO2 catalytic oxidation process for anthracene (ANT) degradation in soil. Wang et al. [16] reported that the degradation efficiency of 2-sec-butyl-4,6-dinitrophenol was significantly improved when ClO2 was heterogeneously catalyzed by Al2O3. Our previous work [18] showed that divalent manganese ions could activate ClO2 for PAH removal from contaminated industrial soil, with the principal active species HOCl and •OH. This research showed that the activation of ClO2 by metal or metal oxide should alleviate the drawbacks seen when ClO2 is used alone. However, the present metal catalysts (Mn- or Cu-based catalysts) pose heavy metal pollution risks to soil. Thus, much more environmentally friendly and effective materials need to be explored for the activation of ClO2 for soil PAH degradation.
With the advantages of being low-cost, environmentally harmless, and commercially available, nanoscale zero-valent iron (nFe0) has been broadly applied to reduce contaminants in water or soil by itself [19,20,21,22,23], or along with the activation of persulfate [24,25]. Micro-scale Fe0 has been reported to be an excellent activator for ClO2 activation for p-nitrophenol removal in aqueous solutions [26]. It is known that ClO2 inevitably forms chlorite (ClO2−) by reacting with pollutant components through a one-electron transfer process (Equation (1)) [27]. Terhalle et al. [28] reported that each consumed ClO2 could generate 62% relative ClO2− and 42% relative hypochlorous acid (HOCl) when ClO2 reacts with organic compounds (e.g., phenol). Aguilar et al. reported on the generation of •OH when using soluble Fe(II) and HOCl due to a Fenton-like reaction (Equation (2)) [29,30]. In addition, Fe(IV) was reported to be generated on nFe0 surfaces when using ClO2− as the oxidant (Equation (3)) [31]. Significantly, soluble Fe(II) species, which can be easily generated during nFe0 oxidation, have been shown to effectively react with ClO2− (Equation (4)) [32,33]. Therefore, nFe0 is a promising activator for the activation of ClO2, but the activation mechanisms, the variation of iron species, and the generated reactive oxygen species are unclear. In addition, nFe0 could act like a fertilizer by promoting plant photosynthesis and iron uptake [20,34,35]. However, no known studies have been conducted describing the degradation routes of PAHs and evaluating the biological toxicity of PAH-contaminated soil treated by a ClO2/nFe0 system. Thus, a study concerning the use of nFe0 to activate ClO2 for the removal of PAHs in soil is worth conducting.
C 6 H 6 O + ClO 2 → C 6 H 4 O 2 +   HOCl   +   ClO 2 -
HOCl + Fe 2 + → Fe 3 + + · OH   +   Cl -
≡ Fe ( II ) + ClO 2 - → ≡ Fe IV = O   +   ClO -
ClO 2 - + 4   Fe 2 + + 4   H + → Cl - + 4   Fe 3 + + 2   H 2 O
Herein, the performance of nFe0-activated ClO2 for PAH removal from soil was investigated. Phenanthrene (PHE) was chosen as a typical PAH to prepare spiked soil, because it belongs to the sixteen parent PAHs which have been grouped into the priority pollutants category by the U.S. Environmental Protection Agency [4]. The main purposes were as follows: (1) to optimize the reaction condition for PAH removal in a ClO2/nFe0 system by investigating the effect of the water/soil ratio, ClO2 dosage, ClO2/nFe0 mole ratio and reaction pHs; (2) to clarify the ClO2 activation mechanisms in the ClO2/nFe0 system by analyzing iron species, exploring ClO2 variations, and detecting the reactive oxygen species through ERR analysis and quenching experiments; (3) to evaluate the toxicity of the degradation intermediates of PAHs in the ClO2/nFe0 system by using quantitative structure–activity relationship prediction technology and detecting lettuces’ growth in ClO2/nFe0-treated soil.

2. Materials and Methods

2.1. Chemicals

Chemicals are listed in detail in Text S1. The PHE-spiked soil samples were prepared as follows: grassland soil samples (5~10 cm) were collected from Fengxian campus in the Shanghai Institute of Technology (Shanghai, China) and dried. Table S1 lists the physicochemical characteristics of the soil.

2.2. Preparation of PHE-Spiked Soil, ClO2 and nFe0

The total iron in soil was analyzed using a previously reported method [36]. The soil was grounded and sieved through a 65-mesh screen. The PHE (10 mg) in 100 mL of n-hexane/dichloromethane (1: 1 v/v) solvent was mixed with the soil (100 g) in a large beaker by magnetic stirring. The PHE-spiked soil was aged in the dark to volatilize the solvent, for at least two weeks, and was used for further experiments.
ClO2 was prepared by the sodium chlorite–sulfuric acid method according to previous research [37]. In brief, 250 mL sulfuric acid solution (0.1 M) was added to 500 mL sodium chlorite solution (0.4 M) by using a peristaltic pump. The generated ClO2 gas was taken out of the reaction flask by N2-purging and purified through a saturated NaClO2 solution. The prepared ClO2 solution was put into a 4 °C refrigerator. The exact concentration of the ClO2 solution was detected before each use [38].
nFe0 was prepared by using NaBH4 to reduce FeSO4 into Fe0 in a liquid phase [39]. Deionized water (100 mL) was added into a 250 mL three-necked flask with N2 was purged for 0.5 h. Then, FeSO4·7H2O (8.00 mg) was put into the flask and stirred at 250 rpm. Sodium borohydride (3.96 g) was dissolved in water (100 mL) and added to flask by peristaltic pump at 4.50 mL/min. The suspension was kept purging N2 for 2 h, and then it was filtered. The precipitates were dried in a vacuum oven for 4 h. Then, the dried nFe0 was collected in a centrifugal tube for further use. The surface morphology and species changes of nFe0 were detected by scanning electron microscopy with energy dispersive spectroscopy (SEM-EDS, ZEISS, Oberkochen, Germany), X-Ray Diffractometer (XRD, D8 Advance davinci, Bruker, Ettlingen, Germany) and an XPS instrument (Thermo Scientific Escalab 250Xi, Waltham, MA, USA). The surface area was analyzed with the Brunauer–Emmett–Teller (BET) nitrogen adsorption technique (ASAP 2010 M+C, Micrometitics Inc., Norcross, GA, USA).

2.3. Batch Degradation Experiments

The batch experiments concerning soil PHE degradation were conducted in 20 mL brown bottles with lids at a constant agitation of 250 (±5) rpm at 25 (±1) °C, avoiding light. The desired PHE-containing soil (3 g) was added to the bottle, followed by proportions of water (i.e., 3, 6, 9, 12, 15 g), ClO2 (i.e., 40, 80,160 mM/kg), and nFe0 (i.e., 0.25, 0.33, 0.50, 1.00, 2.50 g/kg). The effect of varied pHs (3.0~9.0) on PAH degradation in the nFe0/ClO2 system and the ClO2 system was studied, since the soil pH usually ranged from about 6.0 to 8.5 [40]. The pH adjustment of the soil suspension was conducted using 0.1 M of HCl or 0.1 M of NaOH. The different experimental conditions in each batch are summarized in Table S2. The pH variations are presented in Table S3. The extraction procedures for residual PAHs followed the USEPA test method 3550B [41,42]. When reaching the predetermined time, the bottles were immediately put into a −20 °C refrigerator to terminate the reaction for 12 h. After freeze-drying, 5 mL of n-hexane–dichloromethane mixed solvent with a volume ratio of 1: 1 was added to each bottle. The bottle was put on a horizontal oscillator at 250 rpm for 1 h at 25 °C. To extract the residual PAHs in the soil, the capped bottle was ultrasonicated at 400 W for 1 h. After that, the extracting liquid was filtrated through a 0.22 μm PTFE filter membrane and the filtrate was collected for analysis. The recovery rate of this extraction method was around 88%. A GC-MS (Shimadzu, Kyoto, Japan, QP-2010 Ultra) with an SH-Rxi-5Sil-MS column (30 m × 0.25 mm × 0.25 μm, 0.25 μm) was used for the detection of PAHs and intermediates. The carrier gas was helium, and the flow rate was 1.0 mL/min. The injection port was kept at 280.0 °C, and the heating procedure was maintained at 60 °C for 1 min, then increased to 200 °C at 10 °C/min, and then increased to 300 °C at 5 °C/min for 8 min. The MS spectra reading, with an electron impact source at 70 eV, an ion source temperature of 230 °C, and analysis were performed in the SIM mode. When identifying the degradation products, the column temperature programmed for degradation product identification was as follows: starting at 50 °C, to be maintained for 1 min, then warmed from 50 °C to 300 °C at a speed of 5 °C/min, and then maintained at a constant temperature of 300 °C for 5 min, after which identification and analysis of PHA intermediates was performed in SCAN mode.

2.4. Detection of Iron Species and ClO2 Variations

A UV-vis spectrometry device (UV-5500, Metash, Shanghai, China) was employed to analysis iron species at 510 nm [43]. To put it simply, the total Fe(II), including soluble Fe (II) [Fe(II)sol] and surface Fe(II) [Fe(II)surf], was measured by adding 1,10-phenanthroline before filtration, while Fe(II)sol was determined by adding 1,10-phenanthroline after filtration. The concentration of Fe (II)surf was obtained by calculating the differentials between total Fe (II) and Fe(II)sol. The concentration of ClO2 was detected by using UV-vis method at 359 nm and ClO2- was detected at 260 nm after removing ClO2 with N2-blowing [44].

2.5. Identification of Active Oxidant Species

The active oxidant species in the reaction system were identified by using 5,5-dimethyl-1-pyrroline-oxide (DMPO) as a spin trapper with a Bruker micro-ESR device (Standard V2.0, Bruker, Ettlingen, Germany). The resonance frequency was 9.704 G, and the scan time and scan number were 15.5 s and 32, respectively [45]. Methyl phenyl sulfoxide (PMSO), Isopropyl alcohol (IPA), and glycine (Gly) were used as scavengers for Fe(IV), hydroxyl radical (•OH), and hypochlorous acid (HOCl), respectively [14]. The quenching experiments were conducted in PHE-polluted water (1 mg/L) to exclude interferences by soil constitution [11,28,46]. The scavengers were added before oxidant addition, with a scavenger/ClO2 molar ratio of 5:1. As iron minerals in soil mainly existed in the form of hematite (Fe2O3) and goethite (FeOOH) [47,48], the reactive oxygen species in the ClO2/Fe2O3 and ClO2/FeOOH systems were detected to evaluate interferences by the original iron species in the soil.
According to previous studies, Fe(IV) oxidizes PMSO to produce methyl phenyl sulfone (PMSO2) (Equation (5)), while free radicals can oxidize PMSO to produce hydroxylation products (Equation (6)) [49,50,51]. Therefore, PMSO and PMSO2 were, respectively, detected, to examine the contribution of Fe(IV) in the reaction system by using HPLC (Shimadzu, Japan) with an SPD-M20A detector. Text S2 listed the detailed parameters of HPLC, and in Figure S1 the chromatograms of PMSO and PMSO2 are presented.
C 7 H 8 O S + Fe Ⅳ = O   → C 7 H 8 O 2 S + Fe ( Ⅱ )  
C 7 H 8 O S + · OH   → C 6 H 6 O 2 S + · CH 3

2.6. Biological Toxicity Evaluation

The growth of lettuce in PHE-spiked soil treated by ClO2/nFe0 or ClO2 alone were explored, respectively. Lettuce seeds were evenly put on filter paper in a Petri dish for seed germination, and the filter paper was moistened with deionized water. The lettuce seedlings were obtained by incubating the Petri dish at room temperature for 2 days. ClO2/nFe0-treated soil or ClO2-treated soil were used to study lettuce growth, respectively. A total of 100 g of treated soil was placed into each of three pots, with one lettuce seedling planted in each pot. The pots were incubated under light for 16 h and in the dark for 8 h, at room temperature [52]. Water (2 mL) was poured into each pot every day. The plant length, width, and dry weight of lettuce leaves were measured at the selected time. The pots filled with un-contaminated grassland soil were set as the experiment control (CK).

2.7. Data Analysis

Each set of treatments was repeated three times. Data analysis and plotting were performed using origin 2017 (OriginLab, Northampton, MA, USA). The significance of differences between parallel experiments were assessed by using the minimum significance difference postmortem test and univariate with SPSS 24.0 software (SPSS, Inc., Chicago, IL, USA) at a significance level of p < 0.05.

3. Results and Discussion

3.1. Optimization of the Reaction Conditions for PHE Degradation in ClO2/nFe0 System

Measurement of the effect of the water/soil ratio ranging from 1:1 to 5:1 on PHE degradation was conducted in the ClO2/nFe0 system and the ClO2 system, respectively (Figure 1a). As the water/soil ratio increases from 1:1 to 5:1, the degradation of PHE firstly increased by 27.9% (from 30.0% to 57.9%), then decreased by 35.5% (from 57.9% to 22.4%), with the highest degradation of 57.9% occurring at a water/soil ratio of 3:1 in the ClO2/nFe0 system. The reason could be that the increasing of soil moisture to a proper proportion (e.g., 3:1) made part of the pollutants on the soil surface quickly disperse into the aqueous solution. Thus, the contact between pollutants and reactive oxidant species enhanced greatly, with more pollutants desorbed from soil to water at a higher water/soil ratio [53,54]. However, when the water/soil ratio surpassed 3:1, the reactive oxidant species may be consumed by dissolved organic matter before they react with the target pollutants, which was in keeping with previous research [55,56]. In addition, Figure 1a also showed that the removal efficiency of PHE in a ClO2/nFe0 system was higher (5.2~8.7%) than that in the ClO2-alone system, at different water/soil ratios (1:1 except). Based on the above results, a water/soil ratio of 3:1 was recommended for the following experiments.
Pre-experiments showed that the cumulative degradation of PAHs may be negligible, since the PAH residuals were around 93.0–98.0% after 12 h of treatment in the nFe0-alone system. Figure 1b showed the degradation of PHE in soil with varied dosages of ClO2 and a fixed nFe0 dosage (0.33 mg/kg). The residual PAHs significantly decreased from 64.3% to 39.4% as the ClO2 dosage increased from 40 to 160 mM/kg. When the dosage of ClO2 was 40 mM/kg, the stoichiometric efficiency (or the utilization efficiency), defined as the amount of PHE (g) transformed per gram of ClO2 consumed, was 4.3. The utilization efficiency decreased from 3.3 to 1.8 as the ClO2 dosage increased from 80 to 160 mM/kg. The ideal dosage of ClO2 was selected as 80 mM/kg soil, based on the results for degradation efficiency and utilization efficiency.
Different dosages of nFe0 were tested to investigate its effect on the degradation of PHE in the presence of 80 mM/kg of ClO2 (Figure 1c). When the dosage of nFe0 was less than 0.33 g/kg, the degradation of PHE in ClO2/nFe0 system was apparently higher than that in ClO2-alone system. When the nFe0 dosage exceeded 0.33 g/kg, the degradation of PHE decreased. The presence of excess nFe0 could result in the generation of excess Fe(II) species (Equation (7)), which may act as a scavenger to ClO2 and ClO2− via parasitic reactions (Equations (4) and (7)) [32,57,58]. The reduced ClO2 and ClO2− further decreased the generation of reactive oxygen species. Thus, the degradation of PHE decreased with the excess in nFe0. Overall, based on the degradation performance of PHE, the best dosage of nFe0 was determined to be 0.33 g/kg with the molar ratio of ClO2/nFe0 of 15:1.
ClO 2 + 5   Fe 2 + + 4   H + → Cl - + 5   Fe 3 + + 2   H 2 O
The effect of different initial pHs (3.0, 4.5, 6.0, and 9.0) on the degradation of PAHs was shown in Figure 1d. As the pH increased from 3.0 to 9.0, the degradation efficiency of PHE reduced from 57.5% to 18.0% in the ClO2/nFe0 system, and from 50.8% to 7.0% in the ClO2 system. This indicated that acidic conditions favored PHE degradation in the two systems. The reason could be that the increase in pH improved the precipitation tendency for the release of Fe(II) and reduced the generation of reactive oxide species [59,60,61]. Figure 1d also indicated that the presence of nFe0 in ClO2 system increased the degradation efficiency of PHE by approximately 6.5%, 5.2%, and 10.9% at a pH of 3.0, 4.5, and 9.0, respectively. However, when the pH was nearly neutral (pH = 6.0), the removal of PHE in the ClO2/nFe0 system (34.1%) was slightly lower (4.4%) than that in the ClO2-alone system (38.5%). This could be attributed to the reactive oxidant species in the ClO2/nFe0 system, which were much more sensitive to the decline of pH in comparison with those in the ClO2 system. Nevertheless, future experiments are required to determine the changes in reactive oxidant species with varied pHs. The above results indicated that the ClO2/nFe0 system showed superiority over ClO2 for PHE removal at varied pHs. In comparison to other advanced oxidant process-based strategies (Table S4), ClO2/nFe0, for soil PAHs degradation, has a major advantage in high-oxidant-utilization efficiency, with a small oxidant/PAH ratio (w/w).

3.2. Characterization of nFe0

The fresh and used nFe0 were characterized by SEM-EDS (Figure 2). The nFe0 was a spherical particle with a diameter of approximately 25 nm and a BET surface of 73.46 m2/g. It remained spherical after oxidization by ClO2. The surface contains 33.7% Fe and 49.1% O. After the reaction, the composition of Fe decreased to 31.2% and O increased to 53.7%. In addition, Cl (0.1%) was also detected after the reaction. The variance in elemental constituents indicated oxidation of Fe0 by ClO2 (Figure 2f). Figure 3a showed the XRD patterns of original and used nFe0. Two Fe0 characteristic peaks could be found both in fresh and used Fe0 patterns at 2θ values of 44.13° and 64.48°, respectively, indicating that the Fe0 was the main structural composition of the fresh nFe0 (COD2021 standard card 96-901-3472) [62]. The diffraction peaks (30.05°, 35.39°, 43.01°, 56.88°, and 62.46°) of Fe3O4 (maghemite, COD2021 standard card 96-900-5842) or Fe2O3 (magnetite, COD2021 standard card 96-152-8612) appeared in the fresh nFe0, indicating the partial oxidation of fresh nFe0 [63,64]. After the reaction, the characteristic peaks of Fe0 were weaker, but the characteristic peaks of magnetite/maghemite were stronger than those of the fresh nFe0. This indicated that Fe0 was consumed and iron oxide was generated in the ClO2/nFe0 oxidation system.
The surface elemental composition and chemical state of fresh and used nFe0 were characterized by XPS (Figure 3b,c). According to the Fe 2p spectrum, 706.78, 710.25, and 712.61 eV were, respectively, attributed to Fe(0), Fe(Ⅱ), and Fe(Ⅲ) [65]. After participating in degradation, the characteristic peak of Fe(0) disappeared. However, the composition of Fe(Ⅱ) increased from 42.22% to 53.35%, and that of Fe(Ⅲ) increased from 44.61% to 46.65%. This indicated electron transfer in the nFe0, and the formation of iron oxide on the surface. According to the O 1s spectrum (Figure S2), the peaks at 529.5 and 531.0 eV corresponded to the lattice oxygen (Olatt), Fe-O, and surface-absorbed oxygen (Oabs) [66]. After participating in the reaction, the surface lattice oxygen (Olatt) Fe-O content of nFe0 increased from 46.72% to 53.36%, indicating the oxidation of Fe0, while the Oabs gradually decreased from 53.28% to 46.64%, indicating that during the ClO2/nFe0 process, the formation of an oxide layer on the nFe0 surface occurred.

3.3. Reactive Oxygen Species in ClO2/nFe0 System

As a powerful one-electron oxidant, ClO2 could obtain an electron from the pollutant and be turned into ClO2− (Equation (1)) [27]. To explore the intrinsic reaction, the concentrations of ClO2 and ClO2− (Figure 4a) were detected during ClO2 activation by nFe0. The ClO2 concentration in the ClO2/nFe0 system is lower than that in the ClO2 system during the reaction, indicating that much more ClO2 was consumed in the presence of nFe0. For example, the consumed ClO2 in the ClO2/nFe0 system was 11.50 and 76.40 mg/L in the first 5 min and 30 min, but only 5.50 and 41.20 mg/L in the ClO2-alone system, indicating that the presence of Fe0 made the ClO2 be almost doubly consumed. However, the generated ClO2− in the first 5 min was 2.63 mg/L in the ClO2/nFe0 system, which is comparable to that in the ClO2/nFe0 system (2.59 mg/L). At 30 min, the ClO2− in ClO2/nFe0 (2.10 mg/L) was 0.77-fold that of the ClO2 system (2.71 mg/L), respectively. This indicated that the presence of nFe0 did not significantly contribute to the generation of ClO2−, but significantly promoted the consumption of ClO2. To further clarify the reaction between ClO2 and nFe0, Fe(II)surf and Fe(II)sol were analyzed during the reaction and shown in Figure 4b. Compared to the nFe0-alone system, the presence of ClO2 increased both Fe(II)surf and Fe(II)sol. At 2 min and 10 min, the concentration of Fe(II)surf in ClO2/nFe0 was 2.0-fold and 1.9-fold that in the nFe0 system, and the concentration of Fe(II)sol in ClO2/nFe0 was 2.8-fold and 1.6-fold that in the nFe0 system. In addition, the Fe(II)surf concentration was 3.5-fold and 4.0-fold higher than the Fe(II)sol concentration at 2 min and 10 min in the ClO2/nFe0 system, indicating that Fe(II)surf was the main iron species. Accordingly, the concentration of Fe(III)surf and Fe(III)sol in ClO2/nFe0 was 10-fold and 1.5-fold that in the nFe0 system at 10 min (Figure S3), respectively. The concentration of Fe(III)surf was 5.3-fold higher than Fe(III)sol. The results also showed that nFe0 was mainly turned into Fe(II)surf and Fe(III)surf by the provision of electrons to ClO2, similar to other nFe0-based (e.g., nFe0/persulfate) advanced oxidant processes [34,67]
However, though a higher proportion of ClO2 was consumed in the ClO2/nFe0 system, the generated ClO2− was not accordingly increased. Therefore, other active species should have been generated by electron transfer between ClO2 and nFe0. It has been reported that the Cl–O bond of ClO2– could dissociate and react with electron-donating Fe(II) species on an nFe0 surface to generate Fe(IV) (Equation (3)) [31]. Thus, with the existence of ClO2− and Fe(II)surf on nFe0, Fe(IV) should be generated in the ClO2/nFe0 oxidation system.
The reactive oxygen species were further analyzed by EPR detection (Figure 5b). A seven-line signal attributed to the DMPOX adduct was observed both in the ClO2/nFe0 (Figure 5b) system and the ClO2-alone (Figure S4a) system, which might have been generated by the reaction between DMPO and reactive oxidant species (•OH, HOCl, and Fe(IV) species) [19,68]. No signal was detected in nFe0 (Figure S4a). The DMPOX signal in the ClO2/nFe0 system was much stronger in comparison with that in the ClO2-alone system, which indicated that more active substances were produced in the former system. When Gly was added as a specific scavenger of HOCl, the DMPOX signals instantly vanished in the ClO2-alone system (Figure S4a), suggesting that the primary reactive oxygen species responsible for DMPO oxidation was HOCl in the ClO2-alone system. However, after the scavenging of HOCl by Gly, the typical signals of DMPO-OH were observed in the ClO2/nFe0 system (Figure 5b), suggesting the existence of an •OH radical. With further addition of Gly and IPA to quench both •OH and HOCl, the DMPOX signals appeared again. Interestingly, when PMSO was further added as a specific scavenger for Fe(IV), the DMPOX signals disappeared immediately, suggesting Fe(IV) also exists in the ClO2/nFe0 system. The typical signals of DMPO-OH were not observed in the ClO2/Fe2O3 or the ClO2/FeOOH system (Figure S4b,c), let alone Fe(IV) species in the ClO2 system activated by iron minerals in soil. Thus, the iron minerals in soil cannot activate ClO2 to generate reactive oxidant species, and HOCl, •OH, and Fe(IV) were confirmed to exist only in the ClO2/nFe0 system (Scheme S1).
As HOCl and •OH were commonly detected when organic contaminants were oxidized by ClO2 or activated ClO2 [18,28], IPA and Gly were employed as scavengers for •OH and HOCl, respectively, in this study to explore the active species in a ClO2/nFe0 system (Figure 5). As shown in Figure 5a, compared to that in CK (without any scavenger), the addition of Gly markedly increased the residual PHE in a ClO2-alone system and the ClO2/nFe0 system. This indicated that HOCl is involved in the degradation of PHE in the two oxidation systems. The presence of IPA did not significantly affect the residual percentage of PHE in the ClO2-alone system, but resulted in a 55.0% increase in the residual PHE in the ClO2/nFe0 system. This showed that •OH contributed to PHE degradation in the ClO2/nFe0 system.
To further clarify the contribution of HOCl, •OH, and Fe(IV), PMSO was used as an Fe(IV) prober because Fe(IV) can selectively react with PMSO to generate PMSO2 [19]. As shown in Figure 5c, 12.3 μM of PMSO was consumed and 4.8 μM PMSO2 was generated in the ClO2/nFe0 system without any scavenger. Thus, the contribution percentage (η) of Fe(IV), which was defined as the molar ratio of generated PMSO2 to consumed PMSO, was 34.3% in ClO2/nFe0 system. Since the reaction rate of PMSO with •OH (kPMSO, ·OH = 3.61 × 109 M−1s−1) was matched to that of IPA with •OH (kIPA, ·OH = 3 × 109 M−1s−1), IPA was not a suitable scavenger for distinguishing the respective contribution of •OH and PMSO [51,69,70]. As a selective HOCl scavenger [28], Gly was added to the ClO2/nFe0 system to explore the oxidation contribution of HOCl. With the presence of Gly, the consumed PMSO reduced 52.8% (from 12.3 to 5.8 μM), indicating that HOCl contributed 52.8% to the oxidation system. The generated PMSO2 was 2.4 μM and the corresponding η (34.6%) was comparable to that in the oxidation system without any scavenger (34.3%). This showed that Gly hardly affects the function of Fe(IV). Since •OH, Fe(IV), and HOCl were the three main active oxygen species, the contribution of •OH should have been 12.9% (100%–34.3%–52.8%). Therefore, the respective oxidation contribution of Fe(IV), HOCl, and •OH were 34.3%, 52.8%, and 12.9% in the ClO2/nFe0 system. However, the actual oxidant efficiency of the three main active oxygen species for PHE removal is worthy of research in the future.

3.4. Degradation Intermediates and Ecotoxicity Prediction

Tables S5 and S6 list the degradation intermediates of PHE and ANT in the ClO2/nFe0 system using GC-MS (Figures S5 and S6). The main degradation intermediates of PAHs were chlorinated PAHs and oxygenated PAHs [71]. The main intermediate of PHE was 9,10-phenanthrenedione (P1) in the ClO2/nFe0 system (Figure 6a). However, P1 and 9-chlorophenanthrene (P2) were generated in the ClO2-alone system (Figure 6a and Figure S5). The toxicity estimation software tool was employed to estimate the ecotoxicity of the intermediates [11]. The toxicity of P1 was significantly lower than PHE (Figure 7), while P2 still had high mutagenicity toxicity and bioconcentration factor [72,73]. The results showed that the intermediate of PHE in the ClO2/nFe0 system had lower toxicity than those in the ClO2 system. For the degradation of ANT, anthrone (A1), 9,10-anthracinedione (A2), and benzophenone (A3) were detected in both the ClO2/nFe0 and ClO2-alone systems. However, 4-chlorobenzophenone (A4), which was chlorinated from A3, was presented only in the ClO2 system, while anthralin (A5), which takes a dichlorination reaction, existed only in the ClO2/nFe0 system [74,75]. Toxicity estimation analysis (Figure S7) showed that the bioconcentration factor was 614.34 for ANT, but 14.2 for A5 in the ClO2/nFe0 system. Above all, the intermediates of the PAHs in the ClO2/nFe0 system had lower toxicity, indicating a great practical potential of ClO2/nFe0 for soil treatment.

3.5. Lettuce Growth in ClO2/nFe0-Treated Soil

Figure 8 showed the soil-based bioassays of lettuce in soil treated by ClO2/nFe0. With 14 d-cultivation, the lettuce plants grown in ClO2/nFe0-treated soil had longer and wider leaves than those grown in ClO2-treated soil, but were smaller than lettuce grown in clean soil. The dry weights of lettuce in different treatments had no significant differences (Figure S8). This could be attributed to the following reasons: the presence of nFe0 could act as a necessary nutrient for plants [76]. In addition, the intermediates of PAHs in the ClO2/nFe0 system had lower toxicity. However, subsequent studies focusing on enhancing the mineralization of PAHs in soil are still worthy of being carried out.

4. Conclusions

Herein, a nFe0-activated ClO2 system was employed to treat PHE-spiked grassland soil. The reaction conditions (the molar ratio of ClO2 to nFe0, reaction pH, and ratio of water to soil (w/w)) were optimized to improve the degradation efficiency of PHE. By conducting EPR analysis and quenching experiments, Fe(IV), HClO, and •OH were revealed to be the three most important active oxygen species when ClO2 was activated by nFe0 with respective oxidation contributions of 34.3%, 52.8%, and 12.9%. The presence of nFe0 in the ClO2 system decreased the toxicity of degradation intermediates. The lettuces grown in ClO2/nFe0-treated soil displayed better bioassay indexes than those grown in ClO2-treated soil. Overall, these findings support the prospect of remediating organic-pollutant-contaminated soil by using Fe0-activated ClO2 technology.

Supplementary Materials

The following are available online at https://www.mdpi.com/article/10.3390/toxics13010036/s1: Text S1: Materials and reagents; Text S2: Analytical procedures; Figure S1: The chromatograms of PMSO and PMSO2; Figure S2: High-resolution XPS spectra of O 1s for fresh nFe0 (a) and used nFe0 (b); Figure S3: The concentration of Fe(Ⅲ)sol and Fe(Ⅲ)surf in ClO2/nFe0 or nFe0 system. Experimental conditions: [PHE] = 1 mg/L; [ClO2] = 548 mg/L; [nFe0] = 100 mg/L; T = 25 °C; pH = 4.5; Figure S4: Quenching experiments and EPR spectrum of (a) ClO2 system, (b) ClO2/Fe2O3 system and (c) ClO2/FeOOH system under DMPO capture agent. Experimental conditions: [ClO2] = 548 mg/L; [Fe2O3] = 288 mg/L; [FeOOH] = 156 mg/L; [Gly] = 4 mM; [DMPO] = 100 mM; T = 25 °C; reaction time = 10 min; Figure S5: (a)Total ion chromatogram of PHE in different reaction system. The mass spectra of degradation intermediates with retention time at 15.06 min (b) and 17.84 min (c); Figure S6. Total ion chromatogram of ANT in different reaction system (a) and the mass spectra of degradation intermediates with retention time at 17.455 min (b), 17.705 min (c), 13.785 min (d), 15.790 min (e) and 17.515 min (f); Figure S7: Toxicity estimation of ANT and its degradation products in (a,c) ClO2 system and (b,d) ClO2/nFe0 system; Figure S8: Effect of ClO2/nFe0 system on lettuce dry weight; Table S1: The physicochemical properties of the grassland soil; Table S2: Experimental conditions for each batch degradation experiments; Table S3: pH evolution in different treatment after the reaction; Table S4: PAHs degradation efficiencies comparison of ClO2/nFe0 system and other published advanced oxidation techniques; Table S5: The characteristic of PHE and degradation products; Table S6 The characteristic of ANT and degradation products; Scheme S1: proposed mechanisms in ClO2/nFe0 system. References [77,78,79,80] are cited in the Supplementary Materials.

Author Contributions

Conceptualization, X.X., B.L., J.Q., J.M., and X.H.; Methodology, X.X., F.Z., S.C., and J.Q.; Validation, F.Z., S.C., and B.W.; Investigation, X.X., F.Z., and S.C.; Data curation, X.X.; Writing—original draft, X.X.; Writing—review and editing, B.W. and J.Q.; Supervision, J.Q., B.L., J.M., and X.H.; Funding acquisition, J.Q. and X.H. All authors have read and agreed to the published version of the manuscript.

Funding

This research was funded by [the National Natural Science Foundation of China] grant number [No. 42107007] and [the Natural Science Foundation of Shanghai] grant number [No. 23ZR1462300].

Institutional Review Board Statement

Not applicable.

Informed Consent Statement

Not applicable.

Data Availability Statement

The data that support the findings of this study are available from the corresponding author upon reasonable request.

Conflicts of Interest

The authors declare no conflicts of interest.

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Figure 1. Effect of water/soil ratio (a), chlorine dioxide dosage (b), nFe0 dosage (c) and initial pH (d) on the removal of PHE in soil in different reaction systems. Experimental conditions: [PHE]= 100 mg/kg; T = 25 °C; reaction time = 12 h. The different lowercase letters above the bars indicate groups with significant differences (p < 0.05).
Figure 1. Effect of water/soil ratio (a), chlorine dioxide dosage (b), nFe0 dosage (c) and initial pH (d) on the removal of PHE in soil in different reaction systems. Experimental conditions: [PHE]= 100 mg/kg; T = 25 °C; reaction time = 12 h. The different lowercase letters above the bars indicate groups with significant differences (p < 0.05).
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Figure 2. SEM-EDS analysis. SEM image of fresh nFe0 (a) and used nFe0 (d); EDS elemental mapping of fresh nFe0 (b) and used nFe0 (e); and the EDS spectrum of fresh nFe0 (c) and used nFe0 (f). The inset table shows the elemental composition.
Figure 2. SEM-EDS analysis. SEM image of fresh nFe0 (a) and used nFe0 (d); EDS elemental mapping of fresh nFe0 (b) and used nFe0 (e); and the EDS spectrum of fresh nFe0 (c) and used nFe0 (f). The inset table shows the elemental composition.
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Figure 3. XRD spectra (a) and high-resolution XPS spectra of Fe 2p spectra for fresh nFe0 (b) and used nFe0 (c).
Figure 3. XRD spectra (a) and high-resolution XPS spectra of Fe 2p spectra for fresh nFe0 (b) and used nFe0 (c).
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Figure 4. (a) The concentration of ClO2 and ClO2− in ClO2/nFe0 and ClO2 system. (b) The concentration of Fe(II)sol and Fe(II)surf in ClO2/nFe0 or nFe0 systems. Experimental conditions: [PHE] = 1 mg/L; [ClO2] = 548 mg/L; [nFe0] = 100 mg/L; T = 25 °C; pH = 4.5.
Figure 4. (a) The concentration of ClO2 and ClO2− in ClO2/nFe0 and ClO2 system. (b) The concentration of Fe(II)sol and Fe(II)surf in ClO2/nFe0 or nFe0 systems. Experimental conditions: [PHE] = 1 mg/L; [ClO2] = 548 mg/L; [nFe0] = 100 mg/L; T = 25 °C; pH = 4.5.
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Figure 5. Quenching experiments and EPR spectra of ClO2/nFe0 system under DMPO capture agent. (a) Comparison of the quencher effects on the degradation of PHE in ClO2/nFe0 system and ClO2 system. (b) EPR spectra of DMPOX adducts with different scavengers in ClO2/nFe0 system. (c) Effect of Gly on the generation of PMSO and generation of PMSO2 at ClO2/nFe0 system. Experimental conditions: [PHE] = 1 mg/L; [ClO2] = 548 mg/L; [nFe0] = 100 mg/L; [Gly] = [IPA] = [PMSO] = 4 mM; [DMPO] = 100 mM; T = 25 °C; reaction time = 10 min. The different lowercase letters above the bars indicate groups with significant differences (p < 0.05).
Figure 5. Quenching experiments and EPR spectra of ClO2/nFe0 system under DMPO capture agent. (a) Comparison of the quencher effects on the degradation of PHE in ClO2/nFe0 system and ClO2 system. (b) EPR spectra of DMPOX adducts with different scavengers in ClO2/nFe0 system. (c) Effect of Gly on the generation of PMSO and generation of PMSO2 at ClO2/nFe0 system. Experimental conditions: [PHE] = 1 mg/L; [ClO2] = 548 mg/L; [nFe0] = 100 mg/L; [Gly] = [IPA] = [PMSO] = 4 mM; [DMPO] = 100 mM; T = 25 °C; reaction time = 10 min. The different lowercase letters above the bars indicate groups with significant differences (p < 0.05).
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Figure 6. Proposed degradation pathways of PHE (a) and ANT (b).
Figure 6. Proposed degradation pathways of PHE (a) and ANT (b).
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Figure 7. Toxicity estimation of PHE and its degradation products in ClO2 system (a,c,e,g) and ClO2/nFe0 system (b,d,f,h). N/A means not available.
Figure 7. Toxicity estimation of PHE and its degradation products in ClO2 system (a,c,e,g) and ClO2/nFe0 system (b,d,f,h). N/A means not available.
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Figure 8. The leaf length and leaf width of lettuce seedings grown in soil treated with ClO2/nFe0 or nFe0-alone systems. The length of the first leaf (a), the second leaf (b), and the longest leaf (c), and the width of the longest leaf (d). The different lowercase letters above the bars indicate groups with significant differences (p < 0.05).
Figure 8. The leaf length and leaf width of lettuce seedings grown in soil treated with ClO2/nFe0 or nFe0-alone systems. The length of the first leaf (a), the second leaf (b), and the longest leaf (c), and the width of the longest leaf (d). The different lowercase letters above the bars indicate groups with significant differences (p < 0.05).
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MDPI and ACS Style

Hu, X.; Xing, X.; Zhang, F.; Li, B.; Chen, S.; Wang, B.; Qin, J.; Miao, J. Activation of ClO2 by Nanoscale Zero-Valent Iron for Efficient Soil Polycyclic Aromatic Hydrocarbon Degradation: New Insight into the Relative Contribution of Fe(IV) and Hydroxyl Radicals. Toxics 2025, 13, 36. https://doi.org/10.3390/toxics13010036

AMA Style

Hu X, Xing X, Zhang F, Li B, Chen S, Wang B, Qin J, Miao J. Activation of ClO2 by Nanoscale Zero-Valent Iron for Efficient Soil Polycyclic Aromatic Hydrocarbon Degradation: New Insight into the Relative Contribution of Fe(IV) and Hydroxyl Radicals. Toxics. 2025; 13(1):36. https://doi.org/10.3390/toxics13010036

Chicago/Turabian Style

Hu, Xiaojun, Xiaorong Xing, Fan Zhang, Bingzhi Li, Senlin Chen, Bo Wang, Jiaolong Qin, and Jie Miao. 2025. "Activation of ClO2 by Nanoscale Zero-Valent Iron for Efficient Soil Polycyclic Aromatic Hydrocarbon Degradation: New Insight into the Relative Contribution of Fe(IV) and Hydroxyl Radicals" Toxics 13, no. 1: 36. https://doi.org/10.3390/toxics13010036

APA Style

Hu, X., Xing, X., Zhang, F., Li, B., Chen, S., Wang, B., Qin, J., & Miao, J. (2025). Activation of ClO2 by Nanoscale Zero-Valent Iron for Efficient Soil Polycyclic Aromatic Hydrocarbon Degradation: New Insight into the Relative Contribution of Fe(IV) and Hydroxyl Radicals. Toxics, 13(1), 36. https://doi.org/10.3390/toxics13010036

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