Synergistic Remediation of Cr(VI) and P-Nitrophenol Co-Contaminated Soil Using Metal-/Non-Metal-Doped nZVI Catalysts with High Dispersion in the Presence of Persulfate

Abstract

In this work, two novel nanoscale zero-valent iron (nZVI) composites (nanoscale zero-valent iron and copper-intercalated montmorillonite (MMT-nFe⁰/Cu⁰) and carbon microsphere-supported sulfurized nanoscale zero-valent iron (CMS@S-nFe⁰)) were used to treat soil contaminated with both Cr(VI) and p-nitrophenol (PNP), and added persulfate (PMS). Experiments found that the pollutant removal effect has a great relationship with the ratio of water to soil, the amount of catalyst, the amount of PMS, and the pH value. When the conditions are adjusted to the best (water–soil = 2:1, catalyst 30 g/kg, PMS 15 g/kg, pH 7–9), both materials fix Cr(VI) well and decompose PNP. The removal rates of Cr(VI) and PNP by the MMT-nFe⁰/Cu⁰ system are 90.4% and 72.6%, respectively, while the CMS@ S-nFe⁰ system is even more severe, reaching 94.8% and 81.3%. Soil column leaching experiments also proved that the fixation effect of Cr can last for a long time and PNP can be effectively decomposed. Through detection methods such as X-ray diffraction (XRD), Fourier transform infrared spectroscopy (FT-IR) and X-ray photoelectron spectroscopy (XPS), we found that Cr(VI) was effectively reduced to Cr(III) by Fe⁰ and Fe²⁺ ions and subsequently transformed into stable FeCr₂O₄ spinel oxides, and the groups produced after the decomposition of PNP could also help fix the metal. This work provides a way to simultaneously treat Cr(VI) and PNP pollution, and also allows the use of multifunctional nZVI composites in complex soil environments.

1. Introduction

The problem of combined soil pollution is becoming more and more serious, mainly manifested in the joint pollution of heavy metals [1,2,3] (such as Cd, As, and Pb) and organic pollutants [4] (including polycyclic aromatic hydrocarbons and pesticide residues), which harm ecosystems, agricultural product safety, and human health [4,5]. Among these pollutants, hexavalent chromium (Cr(VI)) [6] and p-nitrophenol (PNP) [7] are of particular concern because of their widespread distribution and high toxicity [8]. Cr(VI) is highly mobile and is carcinogenic and teratogenic, while PNP is a refractory nitro aromatic compound that may interfere with endocrine and accumulate in organisms [9,10]. Cleaning up soil contaminated with both Cr(VI) and PNP is difficult because they differ in biogeochemical behavior: Cr(VI) needs to be converted to less mobile and toxic Cr(III) through a reduction reaction to be fixed [11], while PNP requires oxidative degradation to be mineralized [12].

Nanoscale zero-valent iron (nZVI) has become a popular method for treating Cr(VI) and other oxidizing pollutants because of its strong reactivity and good reduction effect [13,14,15]. However, conventional nZVI faces significant limitations in practical applications, particularly for co-contamination scenarios. Its inherent instability leads to rapid oxidation and aggregation, drastically reducing reactivity [16,17]. Crucially, nZVI primarily functions as a reductant and is largely ineffective against co-present organic pollutants like PNP, which require oxidative degradation pathways [18,19]. Furthermore, unmodified nZVI particles exhibit poor mobility and distribution in the complex soil matrix [20]. To overcome the limitations of pristine nZVI, particularly its inability to handle combined redox contaminants, our research group has focused on developing stabilized and multifunctional nZVI composites. Our team has developed materials modified by doping and loading of zero-valent iron [21,22]. Specifically, they fixed nFe⁰ on carbon microspheres (CMSs) and then turned it into a sulfide state, thus creating a multi-layered core–shell structure. Carbon microspheres supported sulfurized nanoscale zero-valent iron (CMS@ S-nFe⁰) material, while a Fe-Cu bimetallic system was innovatively incorporated into montmorillonite (MMT) to fabricate the doped and modified nanoscale zero-valent iron and copper-intercalated montmorillonite (MMT-nFe⁰/Cu⁰) composite, which have been shown to effectively enhance material performance and degrade pollutants in aqueous environments. These materials were designed to integrate chemical reduction (for Cr(VI)) with advanced oxidation processes (AOPs [23,24,25], for organics like PNP), leveraging peroxymonosulfate (PMS) activation to generate powerful sulfate radicals (SO₄·⁻, E₀ = 2.5~3.1 V) [26,27,28].

However, our previous investigations have primarily focused on aqueous systems, whereas soils constitute a more complex environment. Soil constituents (e.g., organic matter, clays, oxides) can scavenge reactive species, compete for adsorption sites, and restrict contaminant accessibility [29], while soil properties such as pH, moisture, and porosity markedly influence reaction kinetics and material stability [5]. Consequently, the promising performance of these nZVI composites in water remediation cannot be directly extrapolated to soil applications. This gap underscores the necessity of systematically exploring their degradation performance and mechanisms in soil, thereby providing the basis for the present study.

Therefore, this study aims to systematically compare the degradation performance of materials prepared using two previously developed modification strategies for nZVI. The specific objectives are to (i) optimize degradation conditions, (ii) evaluate the long-term remediation performance, and (iii) elucidate the degradation mechanisms in soil matrices through characterization methods for the treated soil. By extending the application of nZVI-based materials from aqueous systems to soil environments, this work highlights their potential for effective multi-media pollution remediation.

2. Results and Discussions

2.1. Characterization of nFe⁰, MMT-nFe⁰/Cu⁰ and CMS@S-nFe⁰

In Figure 1, we observed the surface structures of nFe⁰, MMT-nFe⁰/Cu⁰, and CMS@ S-nFe⁰ using a scanning electron microscope (SEM). The freshly made nFe⁰ particles stick together obviously (Figure 1a). However, no clear nFe⁰ particles can be seen on the surface of MMT-nFe⁰/Cu⁰ (Figure 1b), indicating that the nanoparticles are hidden in the clay layer. With increasing Cu loading, the clay layers became thinner and more disordered. For CMS@S-nFe⁰ (Figure 1c), the microspheres displayed a rougher spherical surface as the nFe⁰ loading increased, which confirmed the successful FeS modification on the microsphere surface. Compared with MMT-nFe⁰/Cu⁰, the CMS@S-nFe⁰ system provided a more uniform dispersion of nFe⁰ and better structural stability, indicating superior modification efficiency.

The N₂ adsorption–desorption isotherms and pore size distributions for MMT-nFe⁰/Cu⁰ and CMS@S-nFe⁰ are shown in Figure S1. Both samples showed Type IV isotherms and H4-type hysteresis rings [30], indicating that they have a mesoporous structure. For MMT-nFe⁰/Cu⁰, the pore size distribution is mainly between 2.2 and 4.3 nm, and there is also a peak around 2.1–3.4 nm, indicating that adding Cu helps create more pores. In contrast, CMS@S-nFe⁰ displayed smaller mesopores, as indicated by the closure of the hysteresis loop at a relative pressure of p/p₀ ≈ 0.4 [31]. Brunauer–Emmett–Teller (BET) analysis (Table S1) further showed that the specific surface area of CMS@S-nFe⁰ (88.259 m²/g) was significantly higher than that of nFe⁰ (36.376 m²/g), demonstrating improved dispersion of nFe⁰. Meanwhile, the modified sample exhibited a decreased total pore volume but an enlarged average pore size. Despite its smaller total pore volume compared with MMT-nFe⁰/Cu⁰, the larger surface area and pore size of CMS@S-nFe⁰ facilitated more efficient adsorption and pollutant removal. This structure–performance relationship is consistent with the SEM observations.

Figure S2 shows the Fourier transform infrared spectroscopy (FT-IR) spectra of Fe⁰, MMT-nFe⁰/Cu⁰ and CMS@ S-nFe⁰. In MMT-nFe⁰/Cu⁰, the peak at 3624 cm⁻¹ corresponds to the Al-OH group in montmorillonite. The main peaks at 1103 and 1019 cm⁻¹ are related to the stretching vibrations of the external and internal Si-O skeletons, respectively. In addition, Fe-O stretching vibration signals at 459 and 612 cm⁻¹ indicate the presence of iron oxide in MMT-nFe⁰/Cu⁰. The obvious feature of 526 cm⁻¹ in MMT-nFe⁰/Cu⁰ may come from the Cu-O lattice vibration in copper oxide. For CMS@S-nFe⁰, the characteristic functional groups of carbon spheres appear at 1709 cm⁻¹ (C=O), 1624 cm⁻¹ (C=C), and 1296 cm⁻¹ (C-H). Additional peaks at 1020 cm⁻¹ (C-O) and 670 cm⁻¹ (Fe-O) confirmed the presence of iron oxide. The peak of 1020 cm⁻¹ corresponds to S-O stretching vibration, which may be the result of surface oxidation of FeS. In addition, after the FeS layer was loaded on CMS@nFe⁰, the carbon-related CMS signal and the Fe-O peak intensity of nFe⁰ were significantly weakened, indicating that FeS was successfully deposited on the surface of CMS@nFe⁰.

2.2. Effects of Parameters

The ratio of water to soil, the dosage of different materials, the concentration of PMS, and the pH value will have a great impact on the degradation effect. To design the best repair system, these factors must be adjusted well [32]. As shown in Figure 2, at a water–soil ratio of 1:2, the removal efficiencies of Cr(VI) and PNP were 34.74% and 26.06% in the MMT-nFe⁰/Cu⁰ system (Figure 2a), and 28.67% and 33.12% in the CMS@S-nFe⁰ system (Figure 2b), respectively. The limited removal efficiency under this condition was attributed to incomplete soil wetting [33,34]. When the ratio was increased to 1:1, sufficient soil wetting was achieved, resulting in a marked improvement in removal efficiencies of 63.44% (Cr(VI)) and 38.26% (PNP) for MMT-nFe⁰/Cu⁰, and 64.21% (Cr(VI)) and 50.14% (PNP) for CMS@S-nFe⁰. Further increasing the ratio to 2:1 enhanced removal to 75.46% and 42.26% for Cr(VI) and PNP in the MMT-nFe⁰/Cu⁰ system, and 79.46% and 58.54% in the CMS@S-nFe⁰ system, respectively. No significant improvement was observed at higher ratios (5:1 and 10:1), suggesting that a 2:1 ratio provided nearly complete contaminant leaching and further increase did not enhance removal efficiency. Notably, both systems exhibited similar trends in Cr(VI) immobilization and PNP degradation.

Figure 3 illustrates the removal efficiencies of Cr(VI) and PNP at different material dosages. For both systems, the removal efficiencies of Cr(VI) and PNP increased with increasing dosage. In the MMT-nFe⁰/Cu⁰ system (Figure 3a,b), after 8 h of treatment, the removal efficiencies were 28.67% and 37.76% at dosages of 10 and 20 g/kg, respectively. When the dosage was raised to 30 g/kg, the efficiency sharply increased to 74.63%. Further increases to 40 and 50 g/kg resulted in only slight improvements, with efficiencies of 75.34% and 80.42%, respectively. The most pronounced enhancement occurred between 20 and 30 g/kg, where the removal efficiency nearly doubled (from 37.76% to 74.63%). Beyond 30 g/kg, however, only marginal gains were observed (from 74.63% to 80.42%). The CMS@S-nFe⁰ system (Figure 3c,d) displayed a similar trend but achieved slightly higher Cr(VI) removal efficiency. Considering both cost reduction and the potential adverse impacts of overdosing on the soil environment, 30 g/kg was identified as the optimal application rate.

The effect of MMT-nFe⁰/Cu⁰ and CMS@S-nFe⁰ reaction systems on the removal of Cr(VI) and PNP from the soil was investigated at different PMS concentrations, and the results are shown in Figure 4. As shown in Figure 4a,c, without the addition of PMS, MMT-nFe⁰/Cu⁰ and CMS@S-nFe⁰ had the lowest removal rates for Cr(VI), at 58.67% and 65.48%, respectively. The introduction of 5 g/kg PMS significantly enhanced removal to 77.56% and 85.75%, respectively. However, further PMS increases (5~20 g/kg) maintained stable removal efficiencies within 76–78% and 85–88%, demonstrating that adding a small amount of PMS can promote the removal of Cr(VI), but increasing the concentration of PMS has no significant effect on Cr(VI) removal. PNP removal exhibited distinct concentration dependence (Figure 4b,d). Both systems showed progressive efficiency improvements with increasing PMS concentration (0~15 g/kg). This is mainly due to the increase in SO₄⁻· and ·OH produced by activated PMS, which helps in the oxidation and removal of PNP [35]. Notably, CMS@S-nFe⁰ demonstrated superior activation capability throughout the concentration range. However, when the concentration of PMS increased from 15 g/kg to 20 g/kg, the removal rate of PNP decreased slightly. This phenomenon may be due to the fact that when the concentration of PMS exceeds 15 g/kg, the free radical scavenging ability of PMS increases [36].

As pH significantly influences reaction activity in soil systems, its effects on Cr(VI) and PNP removal by MMT-nFe⁰/Cu⁰ and CMS@S-nFe⁰ were investigated. As shown in Figure 5a, Cr(VI) removal by MMT-nFe⁰/Cu⁰ increased from 51.35% to 80.53% as pH rose from 3 to 9 but declined to 62.45% at pH 11. A similar trend was observed for PNP removal (Figure 5b). In the CMS@S-nFe⁰ system, Cr(VI) removal increased from 55.35% to 92.02% when pH increased from 3 to 7 but decreased to 63.43% at pH 11 (Figure 5c), with PNP removal following the same pattern (Figure 5d). The soil’s acidity and alkalinity strongly influenced pollutant removal. Under acidic conditions (low pH), abundant H⁺ promoted nFe⁰ corrosion, generating Fe(II) that facilitated Cr(VI) reduction and activated PMS to produce SO₄·⁻, enhancing PNP oxidative degradation [37]. Conversely, at higher pH, free and surface-bound Fe(II) was more readily oxidized, forming a dense iron hydroxide layer that hindered nFe⁰ corrosion, reducing Cr(VI) removal [38]. Additionally, in alkaline solutions, SO₄·⁻ reacted with H₂O or OH⁻ to form less reactive ·OH while SO₄·⁻ was consumed, further decreasing PNP removal efficiency [39].

2.3. Degradation Performance in Soil Column

To assess the long-term stabilization of Cr(VI) and PNP in soil column, leaching experiments were performed. As illustrated in Figure 6a, distinct differences in Cr(VI) concentrations were observed in control group, MMT-nFe⁰/Cu⁰, and CMS@S-nFe⁰ systems. At an effluent volume of 50 mL, Cr(VI) concentrations peaked at 17.46 mg/L in the control, 11.4 mg/L in the MMT-nFe⁰/Cu⁰ system, and 7.32 mg/L in the CMS@S-nFe⁰ system. With increasing elution, Cr(VI) concentrations progressively declined, yet the control group exhibited the slowest reduction, remaining as high as 6.02 mg/L at 800 mL. In contrast, the MMT-nFe⁰/Cu⁰ and CMS@S-nFe⁰ systems achieved nearly complete immobilization, with terminal concentrations of 0.06 mg/L and 0.15 mg/L, respectively. These results demonstrate that, although soil columns inherently retain a portion of Cr(VI) through physical filtration, the introduction of modified materials substantially enhances Cr(VI) fixation, with CMS@S-nFe⁰ exhibiting superior stabilization efficacy compared to MMT-nFe⁰/Cu⁰.

Similarly, Figure 6b depicts the PNP leaching behavior across the three systems. At 100 mL, the maximum PNP concentrations in the effluent were 9.16 mg/L in the control, 6.20 mg/L in the MMT-nFe⁰/Cu⁰ system, and 5.62 mg/L in the CMS@S-nFe⁰ system. Although both amendments facilitated PNP removal compared with the control, the removal efficiency was lower than that observed for Cr(VI), likely due to the weaker affinity of PNP for soil surfaces. During continuous leaching, PNP concentrations gradually decreased, with the control group again exhibiting the slowest decline (2.34 mg/L at 800 mL). In contrast, the MMT-nFe⁰/Cu⁰ and CMS@S-nFe⁰ systems achieved near-complete PNP removal by 500 mL, stabilizing at 0.06 mg/L and 0.05 mg/L, respectively. Notably, the CMS@S-nFe⁰ system consistently outperformed MMT-nFe⁰/Cu⁰, reflecting the higher density of reactive sites and stronger adsorption/complexation capacity of CMS, which promoted more efficient PNP stabilization.

To further elucidate the transformation of Cr speciation in different systems, soil samples were taken from the bottom layer of the columns after the leaching experiments, and the five fractions of Cr were determined using the Tessier sequential extraction method [40].The results (Figure 7) reveal distinct differences in morphological distribution of Cr among the three systems (control, MMT-nFe⁰/Cu⁰, and CMS@S-nFe⁰). Compared with the control group, the contents of exchangeable, carbonate-bound, ferromanganese oxide and residue-bound states of heavy metal Cr in soil were significantly reduced, while the contents of organic-bound states were elevated in the MMT-nFe⁰/Cu⁰ and CMS@S-nFe⁰ reaction systems. The organic-bound form became the dominant Cr species, followed by the residual fraction. Given that the mobility of Cr fractions follows the order exchangeable > carbonate-bound > organic-bound > residual, these findings indicate that the addition of the material effectively converts Cr from an unstable form that is relatively easy to leach to a more stable form [41].

The leaching concentrations of metals in three sets of soil columns were evaluated using the TCLP leaching toxicity test [42] to assess the potential environmental risks of secondary metal release during remediation. As shown in Table S2, compared with the MMT-nFe⁰/Cu⁰ system, the CMS@S-nFe⁰ system exhibited a higher total Fe leaching concentration but a lower total Cr concentration. This phenomenon may be attributed to the stronger corrosive effect of the CMS@S-nFe⁰ system on the Fe surface of the catalyst during the reaction, whereas the adsorption capacity of MMT effectively inhibited the dissolution of Fe ions in the nFe⁰ system [43,44]. In contrast, the CMS@S-nFe⁰ system demonstrated stronger immobilization of Cr, which further confirmed that the Cr(VI) removal efficiency of CMS@S-nFe⁰ was superior to that of MMT-nFe⁰/Cu⁰.

2.4. Remediation Mechanism

To elucidate the reaction mechanisms and stabilization pathways of Cr(VI) in the treated soils, a series of characterization analyses, including X-ray diffractometer (XRD), FT-IR, and X-ray photoelectron spectroscopy (XPS), were performed. The XRD patterns of soil samples before and after treatment are presented in Figure 8a. The MMT-nFe⁰/Cu⁰ treated soil showed no new characteristic peaks beyond the original SiO₂ peaks, likely due to the low heavy metal content. In contrast, distinct FeCr₂O₄ peaks [16] were observed in the CMS@S- nFe⁰ system, indicating effective Cr(VI) reduction to Cr(III) and subsequent formation of Fe-Cr oxides [45]. This transformation suggests that Fe⁰ served as an electron donor, promoting the conversion of soluble Cr(VI) to stable FeCr₂O₄ spinel oxides through redox coupling reactions between Fe and Cr species [46,47].

FT-IR spectra of the treated soils (Figure 8b) exhibited a characteristic absorption band observed at approximately 497 cm⁻¹, attributed to the Fe-O stretching vibration of iron oxides (Fe₂O₃/Fe₃O₄) formed during the redox transformation of Fe⁰, confirming the presence of iron oxide species bonded with the soil matrix with markedly enhanced intensities in both MMT-nFe⁰/Cu⁰ and CMS@S-nFe⁰ systems compared to untreated soil. The enhanced intensities at 3480, 1624, 1442, and 1015 cm⁻¹ correspond to N-H, C=O, C-N, and C-O stretching vibrations [48], respectively, which can be assigned to the formation and adsorption of carbon-containing and nitrogen-containing intermediates such as p-aminophenol, hydroquinone, benzenetriol, and 4-nitrocatechol generated during PNP degradation [49]. These intermediates introduced additional hydroxyl, amino, and carbonyl functional groups on the soil surface, thereby strengthening metal–organic complexation [50] and Cr(VI) immobilization. Notably, CMS-based materials provided more abundant reactive sites and functional moieties than MMT, further facilitating pollutant immobilization.

XPS analysis provided further evidence supporting the proposed Cr(VI) removal mechanism in the CMS@S-nFe⁰ system. The Fe 2p were deconvoluted into four main components corresponding to the Fe 2p_3/2 and Fe 2p_1/2 spin–orbit doublets of Fe²⁺ and Fe(III) species, along with distinct satellite features. As shown in Figure 8c, the Fe 2p_3/2 peaks located at 712.45 eV (Fe(II)) and 714.43 eV (Fe(III)), and a noticeable satellite peak (Sat., 718.56 eV) attributed to Fe(III), together with their corresponding Fe 2p_1/2 peaks at 724.3 eV and 727.06 eV, respectively, confirm the coexistence of mixed-valence iron species [51]. The Fe 2p spectra revealed a notable shift in the Fe valence state before and after the reaction. The proportion of Fe(II) decreased from 67.4% to 51.8%, while Fe(III) increased from 32.6% to 48.2%, indicating the progressive oxidation of Fe(II) to Fe(III) during the reaction. This transformation suggests that Fe⁰ and Fe(II) were actively involved in electron transfer processes, serving as major reductants for Cr(VI) reduction. The concurrent increase in Fe(III) content implies the formation of Fe (hydr)oxides, such as Fe(OH)₃ or FeOOH, which not only act as oxidation products but also provide additional adsorption and co-precipitation sites for Cr species. Similarly, the Cr 2p spectra (Figure 8d) were deconvoluted into the characteristic Cr 2p_3/2 and Cr 2p_1/2 components corresponding to Cr(III) and Cr(VI) species. The fitted peaks appeared at 575.72 eV and 579.85 eV for Cr 2p_3/2, and at 584.08 eV and 588.92 eV for Cr 2p_1/2, respectively [52,53]. Quantitative analysis showed that Cr(III) accounted for approximately 46.87% of the total chromium after reaction, confirming that Cr(VI) was predominantly reduced to Cr(III). The coexistence of Cr(VI) and Cr(III) signals suggests that reduction and surface complexation occurred concurrently, forming stable Fe-O-Cr linkages on the CMS@S-nFe⁰ surface. XPS elemental analysis results are summarized in Table S3.

Figure 9 illustrates the proposed mechanism for the synergistic remediation of Cr(VI) and p-nitrophenol (PNP) in soil using metal-/non-metal-doped nZVI catalysts with high dispersion under persulfate activation. The pH value plays a crucial role in regulating the redox transformation of Cr species and the evolution of Fe phases. Under acidic conditions (pH < 6), abundant protons accelerate Fe⁰ corrosion, leading to the continuous release of Fe(II) and Fe(III) species. These species act as key reductants and intermediates that rapidly reduce Cr(VI) to Cr(III) through Equations (1–3) [54]. Meanwhile, the enhanced oxidation of Fe⁰ promotes the formation of amorphous Fe(OH)₃ or FeOOH, which can adsorb or co-precipitate Cr(III), contributing to its immobilization. In addition, the degradation of PNP generates hydroxyl, amino, and carbonyl functional groups, which facilitate the adsorption and complexation of residual metal ions. As the pH increases toward neutral or slightly alkaline conditions, Fe⁰ corrosion becomes slower, but the stability of secondary Fe and Cr phases improves. Under such conditions, Fe(II) and Cr(III) species tend to transform into more crystalline and thermodynamically stable FeCr₂O₄ spinel oxides according to Equation (4) [55], leading to the long-term structural fixation of Cr(III) and decreased mobility. Compared with MMT-nFe⁰/Cu⁰, the CMS@S-nFe⁰ system provides more reactive sites and facilitates electron transfer, thereby promoting both the formation of Fe-Cr spinel oxides and the activation of persulfate for efficient PNP degradation. Therefore, pH governs both the kinetics and transformation pathways of Cr(VI), where acidic environments favor rapid reduction and amorphous Fe (hydr)oxide formation, while neutral alkaline conditions promote spinel oxide precipitation and persistent Cr stabilization.

$${\mathrm{C}\mathrm{r}}_2{\mathrm{O}}_7^{2-}+3{\mathrm{F}\mathrm{e}}^0+14{\mathrm{H}}^{+}\to 3{\mathrm{F}\mathrm{e}}^{2+}+2{\mathrm{C}\mathrm{r}}^{3+}+7{\mathrm{H}}_2\mathrm{O}$$

(1)

$$\mathrm{H}\mathrm{C}\mathrm{r}{\mathrm{O}}_4^{-}+3{\mathrm{F}\mathrm{e}}^{2+}+{7\mathrm{H}}^{+}\to 3{\mathrm{F}\mathrm{e}}^{3+}+{\mathrm{C}\mathrm{r}}^{3+}+4{\mathrm{H}}_2\mathrm{O}$$

(2)

$$2\mathrm{C}\mathrm{r}{\mathrm{O}}_4^{2-}+3{\mathrm{F}\mathrm{e}}^{2+}+8{\mathrm{H}}^{+}\to 2{\mathrm{C}\mathrm{r}}^{3+}+3{\mathrm{F}\mathrm{e}}^{3+}+4{\mathrm{H}}_2\mathrm{O}$$

(3)

$${\mathrm{C}\mathrm{r}}_2{\mathrm{O}}_7^{2-}+3{\mathrm{F}\mathrm{e}}^0+8{\mathrm{H}}^{+}\to 2\mathrm{F}\mathrm{e}{\mathrm{C}\mathrm{r}}_2{\mathrm{O}}_4+4{\mathrm{H}}_2\mathrm{O}$$

(4)

3. Materials and Methods

3.1. Chemical and Reagents

NaOH (AR), HCl (AR), K₂Cr₂O₇ (AR), C₆H₅NO₃ (AR), MgCl₂·6H₂O (AR), C₂H₃NaO₂ (AR), NH₂OH·HCl (AR), C₂H₄O₂ (GR), C₂H₇NO₂ (AR), HNO₃ (AR), HClO₄ (GR). All are produced by Sinopharm Chemical Reagent Co., Ltd. (Shanghai, China). KHSO₅^.0.5KHSO₄·0.5K₂SO₄ (≥42% KHSO₅) are purchased from Aladdin Biochemical Technology Co., Ltd. (Shanghai, China). All the water used in the experiment was completely deoxidized and deionized water.

3.2. Preparation of MMT-nFe⁰/Cu⁰ and CMS@S-nFe⁰

MMT-nFe⁰/Cu⁰ and CMS@S-nFe⁰ were synthesized following previously established methods [21,22]. The detailed information is presented in the Supplementary Materials (Text S1).

3.3. Soil Preparation

The original soil was collected from the farmland of Jianshe Town, Chongming District, Shanghai, China. The basic physicochemical characteristics of the tested soil, including organic matter content, clay fraction, redox potential, and buffering capacity, are summarized in Table S4. Surface weeds, stones and other foreign substances were removed. After it dried naturally, it was ground through a 2 mm sieve and set aside for later use. The concentration of soil complex pollution was set at 100 mg/kg Cr(VI) and 50 mg/kg PNP. The preparation process was as follows: First, a certain amount of K₂Cr₂O₇ and PNP reagents was weighed to prepare 100 mg/L Cr(VI) and 100 mg/L PNP solutions. Take 1 kg of the reserved raw soil and put it in a 1 L beaker. Pour 1000 mL of Cr(VI) and 500 mL of PNP solution into the soil. Place it under a spiral stirrer and stir for 12 h to mix evenly. Subsequently, the soil was naturally air-dried at ambient temperature and homogenized by grinding to pass through a 1 mm sieve. Then it was placed in a sealed container and kept in a refrigerator at 4 °C for one month before being used as the test soil.

3.4. Batch Experiments

Degradation experiments of Cr(VI) and PNP were carried out in 100 mL conical flasks containing 5 g of composite contaminated soil and deionized water. The influences of water–soil ratio (0.5–10, v:w), catalyst dosage (10–50 g/kg of MMT-nFe⁰/Cu⁰ or CMS@S-nFe⁰), PMS dosage (0–20 g/kg), and pH (3–11) were systematically evaluated. Unless otherwise specified, the reaction conditions were set as follows: water–soil ratio of 2, catalyst dosage of 30 g/kg, PMS dosage of 15 g/kg, and pH = 5. The mixtures were shaken at 200 rpm in a constant-temperature air bath (25 °C) for 8 h. At predetermined intervals, 2 mL aliquots were withdrawn, centrifuged (5000 rpm, 10 min), and filtered through a 0.45 μm membrane prior to the determination of Cr(VI) and PNP concentrations.

3.5. Soil Column Leaching Experiment

The soil column leaching experiments were performed using a purpose-designed soil column apparatus (Figure S3); detailed information on the design of the soil column device is in Text S2. Soil column experiments were conducted with three groups: control, MMT-nFe⁰/Cu⁰, and CMS@S-nFe⁰. Each column was packed with 200 g of soil, and the amendments were mixed with the top 1–2 cm layer together with PMS (15 g/kg). Columns were saturated with 1 L deionized water at a flow rate of 0.15 mL/min and allowed to equilibrate for 120 h. Continuous leaching was then performed at room temperature using a Mariotte bottle under constant head, and effluent samples were collected periodically for Cr(VI) and PNP analysis.

The morphological distribution of Cr in soil was analyzed using the Tessier sequential extraction method, and the leaching toxicity was evaluated by the TCLP procedure. The detailed information is presented in the Supplementary Materials (Text S3 and Text S4).

3.6. Characterization

We used a scanning electron microscope (SEM, S-4800, Hitachi, Tokyo, Japan) to observe the shape characteristics of the samples. The surface area was measured using the Brunauer–Emmett–Teller (BET) method, the equipment was a nitrogen adsorption–desorption instrument (BET, Autosorb-iQ-2MP, Quantachrome Instruments, Boynton Beach, FL, USA), and the pore size distribution was calculated using Barrett–Joyner–Halenda (BJH) analysis. Fourier transform infrared spectroscopy (FTIR, Nicolet iS10, Thermo Fisher Scientific, Waltham, MA, USA) was used to detect functional groups on the sample surface. An X-ray diffractometer (XRD, D8 Advance, Bruker, Karlsruhe, Germany) was used to determine the crystal phase. X-ray photoelectron spectroscopy (XPS, ESCALAB 250Xi, Thermo Fisher Scientific, Waltham, MA, USA) was used to analyze the material composition. Additional information about XRD and XPS can be found in Text S5.

3.7. Analytic Methods

The Cr(VI) concentration was determined by dibenzoyl dihydrazide (C₁₃H₁₄N₄O) spectrophotometric method (GB7467-87). The corresponding absorbance values were measured at 540 nm by taking Cr(VI) standard solution with different concentration gradients. The labeled curves are plotted as follows:

$$A=0.7563C+0.0012 (R^2=0.9999)$$

(5)

where A is the absorbance value; C is the Cr(VI) concentration (mg/L); and R² is the curve correlation coefficient.

The total chromium content was measured using inductively coupled plasma spectroscopy (ICP-MS, NexIon 300X, PerkinElmer Co., Ltd., Waltham, MA, USA). The amount of Cr(III) is calculated by subtracting the content of Cr(VI) from total chromium.

The concentration of PNP was determined at a wavelength of 400 nm using a UV–Vis spectrophotometer (UV-2600, Shimadzu Corporation, Kyoto, Japan). The corresponding absorbance values were measured by taking different concentration gradients of the PNP standard solution (pH ≥ 9), and the labeled curves were plotted as follows:

$$A=0.1274C+0.0017 (R^2=0.9993)$$

(6)

where A is the absorbance value; C is the PNP concentration (mg/L); and R² is the curve correlation coefficient.

The removal rate of Cr(VI) and PNP was calculated as follows:

$$R\left(\%\right)={\displaystyle \frac{C_0-C_t}{C_0}}\times 100\%$$

(7)

where R is the Cr(VI) and PNP removal rate (%); C₀ is the initial concentration of Cr(VI) and PNP (mg/L); and C_t is the remaining concentration of Cr(VI) and PNP after reaction time t (mg/L).

4. Conclusions

This study demonstrated the effective remediation of Cr(VI) and p-nitrophenol (PNP) co-contaminated soil using two modified nZVI composites, MMT-nFe⁰/Cu⁰ and CMS@S-nFe⁰, in the presence of PMS. Batch experiment and soil column leaching experiment confirmed that both materials significantly enhanced pollutant removal, with CMS@S-nFe⁰ consistently exhibiting superior performance due to its higher dispersion, larger surface area, and stronger fixation capacity. Cr(VI) was effectively reduced to Cr(III) by Fe⁰ and Fe²⁺ ions and subsequently transformed into stable FeCr₂O₄ spinel oxides, while PNP underwent oxidative degradation and adsorption. Long-term soil column tests further revealed that labile Cr fractions were transformed into more stable forms, indicating sustained immobilization and reduced leaching risks. Compared with MMT-nFe⁰/Cu⁰, CMS@S-nFe⁰ showed enhanced PMS activation and greater stabilization efficiency, highlighting its potential as a multifunctional amendment for complex soil remediation. Overall, this work provides new insights into the synergistic redox mechanisms of modified nZVI composites and expands their applicability to addressing heavy metal–organic co-contamination in soil systems.