Anti-Interference Fe-N-C/PMS System: Synergistic Radical-Nonradical Pathways Enabled by sp² Carbon and Metal-N Coordination

Abstract

Phenol is a refractory organic pollutant that is difficult to degrade in wastewater treatment, and efficiently and stably degrading phenol presents a significant challenge. In this study, iron-doped humic acid-based nitrogen–carbon materials were prepared to activate peroxymonosulfate (PMS) for the degradation of phenol. The Fe-N-C/PMS system achieved a phenol degradation rate of 99.71%, which follows a first-order kinetic model, with the reaction rate constant of 0.1419 min⁻¹. The phenol degradation rate remained above 92% in inorganic anions (Cl⁻, SO₄²⁻, HCO₃⁻) and humic acid and the system maintained a 100% phenol removal rate over a wide pH range (3–9). The iron in the catalyst predominantly exists in the forms of Fe⁰ and Fe₃C, and Fe⁰, Fe²⁺/Fe³⁺ are the main active sites that promote PMS activation during the reaction. Additionally, Fe-N-C has a large specific surface area (1041.36 m²/g). Quenching experiments and electron spin resonance (ESR) spectroscopy detected the active free radicals in the Fe-N-C/PMS system: SO₄^•−, •OH, O₂^•−, and ¹O₂. The mechanism for phenol degradation was discussed, involving radical pathways (SO₄^•−, •OH, O₂^•−) and the non-radical pathway (¹O₂), in the Fe-N-C/PMS system activated by Fe⁰, Fe²⁺/Fe³⁺, sp² hybridized carbon, C-O/C-N, C=O, and graphitic nitrogen active sites. This study provides new insights into the synthesis of efficient carbon-based catalysts for phenol degradation and water remediation.

1. Introduction

Phenol is one of the difficult-to-degrade organic pollutants, harmful to human health and environment [1]. Phenol can cause severe toxicological effects on aquatic animals and plants in the surrounding water environment, thereby damaging human health [2]. Advanced oxidation processes (AOPs) based on persulfate activation can in situ generate powerful chemical oxidants, such as hydroxyl radicals (•OH), sulfate radicals (SO₄^•−), and superoxide anion radicals (O₂^•−). These free radicals can degrade organic pollutants [3,4,5].

In recent years, researchers have explored various types of materials for organic pollutants degradation through radical and/or non-radical mechanisms, such as activated carbon [6], carbon nanotubes [7], graphene [8], and reduced graphene oxide [9]. Natural polymer humic acid (HA) is an excellent precursor for carbon materials and is widely present in natural environment. The doping of heteroatoms into carbon-based materials can introduce more active sites [10], significantly enhancing the catalytic activity of PMS/peroxydisulfate(PDS) for organic pollutants degradation. Nitrogen is one of the most widely used heteroatoms for doping. Graphitic carbon nitride (g-C₃N₄) is a compound with high nitrogen content, a graphite-like structure, and is cheap and stable [11]. Using g-C₃N₄ as a nitrogen precursor avoids the need for a subsequent template removal process, and the gases released during the decomposition facilitate the formation of the carbon material’s pore structure, further increasing the specific surface area of the material.

The disadvantage of using transition metals as persulfate catalysts is the potential for metal leaching, which can cause secondary pollution. However, the larger specific surface area and higher stability of carbon materials can disperse and stabilize metal particles, inhibiting metal leaching and promoting electron transfer [12]. Compared to single transition metals and carbon materials, metal-carbon composite catalysts exhibit excellent catalytic performance for persulfate activation, with superior stability [13,14]. Carbon-based iron catalysts show strong activation activity towards persulfate.

The composite structure formed by associating nitrogen-doped carbon-based materials with metal components can endow the catalyst with new physicochemical properties, thereby further enhancing its catalytic ability. Fe-doped nitrogen–carbon materials exhibit excellent performance for organic pollutant degradation over a wide pH range [15,16,17]. Transition metals and carbon can play a synergistic role, forming a core–shell structure with nitrogen–carbon elements, and the metal-nitrogen component serves as the activation active site for persulfates [18,19,20]. Meanwhile, carbon can also weaken the leaching of internal metal nanoparticles in water treatment environment [21,22]. In addition, the metal particles can recover catalyst through magnetic response [23]. Therefore, the development of high-performance iron-nitrogen co-doped carbon materials is of great significance.

For example, Wang et al. reported the degradation of phenol using Fe₃C nanoparticles wrapped in nitrogen-doped carbon nanotubes. The quaternary and pyridine nitrogen atoms showed a synergistic effect on Fe₃C and carbon, with the active sites being iron on the carbon surface [24]. Pang et al. prepared Fe₃O₄@α-Fe₂O₃-N-rGO catalysts, which activated PDS and almost completely removed 2,4-DCP [25]. Long et al. [26] studied the catalytic activity of Fe/N co-doped carbon materials with different structures for the degradation of chlorophenol. Similarly, Xu et al. [27] activated persulfate using Fe/N co-doped biochar to degrade organic pollutants. These studies have demonstrated enhanced catalytic activity of Fe/N co-doped carbon materials. To further increase the number of active sites on the material’s surface and improve its stability, this study also chose to introduce the transition metal (iron) into nitrogen-doped carbon materials [28,29]. This modification can improve the electronic structure, maintaining the high activation ability of the metal while coupling with carbon material to synergistically activate persulfate [30]. Furthermore, the addition of iron imparts magnetic properties to the material, enabling rapid separation from the solution using an external magnetic field, which addresses the challenge of catalyst recovery.

This study prepared iron-nitrogen co-doped carbon materials by high-temperature carbonization using HA as the carbon source, g-C₃N₄ as the nitrogen source, and FeCl₃·6H₂O as the iron source. The effect of iron doping on the surface morphology and structure of the materials was investigated. The impact of factors such as oxidant type, initial pH value, inorganic anions, and natural organic matter on phenol removal performance was systematically studied and compared with N-C. The stability of the materials was evaluated through cyclic reuse experiments. The catalytic reaction mechanism of the activation system was revealed to detect the types of active species in the solution and the structural changes of the catalyst before and after the reaction.

2. Results and Discussion

2.1. Analysis of Precursor Properties

The elemental analysis of HA is shown in

Table 1. Element analysis of HA.

Sample (wt. %)	Cad	Had	Nad	Sad	Oad
HA	46.04	2.77	1.09	0.46	49.64

ad: air-dried basis.

. HA is mainly composed of C, O, H, and small amounts of N and S, with the carbon content reaching 46.04%. Additionally, HA contains trace amounts of metal elements such as iron, calcium, and potassium. These metal elements can catalyze the formation of graphite and disordered layered structures during high-temperature carbonization, providing active sites for the material.

The FTIR analysis results of HA are shown in Figure 1a. The absorption peak near 3440 cm⁻¹ corresponds to the stretching vibration of -OH [31]; the absorption peak near 2916 cm⁻¹ corresponds to the asymmetric stretching vibration of -CH₂- [32]; the absorption peak near 1598 cm⁻¹ corresponds to the asymmetric stretching vibration of C=O in -COO- and -COOH groups [32]; the absorption peak near 1382 cm⁻¹ corresponds to the aromatic ring stretching vibration [32]. These functional groups can influence the structural evolution and catalytic performance of the carbon material during preparation. Thermogravimetric (TG) and the derivative thermogravimetry (DTG) curves of HA are shown in Figure 1b. The weight loss of HA can be divided into two main stages: the weight loss in the range of 50–200 °C was primarily due to moisture removal, while the weight loss in the range of 300–600 °C was mainly caused by the breaking of macromolecular structures and the decomposition of oxygen-containing functional groups. Under a nitrogen atmosphere, during pyrolysis from room temperature to 800 °C, the carbon yield of HA reached 51%, which is higher than most biomass materials. This indicates that HA is a promising precursor for carbon materials.

The FTIR analysis results of g-C₃N₄ are shown in Figure 1c. The peak around 3000–3300 cm⁻¹ corresponds to the stretching vibrations of O-H and N-H [33]; the peak around 1200–1700 cm⁻¹ corresponds to the stretching vibration of C-N in the C₆N₇ ring [32]; the peak near 809 cm⁻¹ corresponds to the breathing vibration of the heptazine ring [34], indicating that g-C₃N₄ was successfully synthesized. Thermogravimetric (TG) and derivative thermogravimetry (DTG) curves of g-C₃N₄ are shown in Figure 1d. g-C₃N₄ underwent only one distinct weight loss stages, with a decomposition temperature of 575 °C, and it completely decomposed at 670 °C. The DTG curve shows a peak around 630 °C, during which a large amount of NH₃, CXNYHZ, and other gases were generated.

The SEM results of g-C₃N₄, as shown in Figure 2, reveal that the surface of g-C₃N₄ contains numerous wrinkled, sheet-like structures. This three-dimensional structure, formed by the curling and winding of two-dimensional sheets, can serve as an excellent template for material synthesis. In conclusion, using g-C₃N₄ as a nitrogen source and a carbonization temperature above 670 °C for the material allows the elimination of the subsequent template removal. Additionally, the nitrogen-containing small molecules released during thermal decomposition can be incorporated as nitrogen dopants into the HA carbon material precursor, thereby enhancing the nitrogen doping level of the material.

2.2. Characterization of Fe-N-C

The SEM images of Fe-N-C at different magnifications are shown in Figure 3. Compared to N-C (Figure S1), the surface of Fe-N-C exhibits a higher degree of disorder and irregularity. The Fe-N-C primarily shows two types of morphologies: one is a loose, porous, coral-like structure with layered folds, which is formed by the high-temperature decomposition of g-C₃N₄ as a template; the other presents a curved, worm-like tubular structures, which is the result of iron acting as a catalyst to promote the formation of tubular structures. The EDS analysis of Fe-N-C is shown in Figure 3c.

As shown in Figure 4a, both Fe-N-C and N-C exhibit type IV isotherms, indicating the presence of mesopores and micropores in Fe-N-C. The pore size distribution of Fe-N-C and N-C is shown in Figure 4b. Fe-N-C has a higher number of micropores below 2 nm and a more uniform pore size distribution in the 2–30 nm range compared to N-C, suggesting that Fe-N-C has a more uniform distribution of pores across different scales. The specific surface areas of Fe-N-C and N-C are 1041.36 and 247.869 m²/g, respectively, and the total pore volumes are 2.172 and 0.676 cm³/g. Compared to N-C, Fe-N-C exhibits a significantly larger specific surface area. These results are consistent with the SEM images, indicating that Fe/N co-doping enhances the specific surface area and pore volume of the carbon-based catalyst, thus promoting the adsorption and transfer of PMS and phenol during the degradation process.

The XRD pattern of Fe-N-C is shown in Figure 4c. Like N-C, Fe-N-C exhibits a diffraction peak at 26°, corresponding to the (002) crystal plane of carbon. However, this peak is sharper in Fe-N-C compared to N-C at 26°, indicating that the doping of iron improves the graphitization degree of the material. In addition, after the doping of iron, several new diffraction peaks appear in the XRD pattern of Fe-N-C. The diffraction peaks at 44.6° and 65.0° correspond to the (110) and (200) crystal planes of Fe⁰ (JCPDS No. 06-0696), respectively. The peaks at 45.0°, 43.7°, 42.9°, 44.6°, 45.9°, 37.6°, 49.1°, and 37.7° correspond to the (031), (102), (211), (220), (112), (121), (221), and (210) crystal planes of Fe₃C (JCPDS No. 35-0772), respectively. These results indicate that iron has been successfully doping into the material in the forms of Fe₀ and Fe₃C. FeCl₃·6H₂O likely plays two roles during the catalyst preparation: one is catalyzing the formation of a graphitic structure in the carbon material [35], and the other is serving as an iron source incorporated into the material.

The Raman spectrum is shown in Figure 4d. Fe-N-C displays peaks at 1340 cm⁻¹ and 1580 cm⁻¹, corresponding to the D and G band. A lower I_D/I_G ratio indicates a higher degree of graphitization. The I_D/I_G ratio of Fe-N-C is 0.65, which is significantly smaller than the value of 0.83 for N-C. Additionally, the peaks for Fe-N-C are sharper compared to N-C, suggesting that doping with iron enhances the graphitization degree of the material within a certain range.

The Fe-N-C sample was analyzed by XPS as shown in Figure 5. The sample exhibits signal peaks near binding energies of C (284.8), N (401.2), O (530.8), and Fe (720 eV). This indicates that Fe and N elements have been incorporated into the material. As shown in Figure 5b, the C1s spectrum is fitted into four characteristic peaks [36]: C=C (284.0 eV), C-C (284.6 eV), C-O/C-N (287.6 eV), and C=O (290.6 eV). As shown in Figure 5c, the high-resolution N1s spectrum is fitted into four characteristic peaks [37]: pyridinic N, pyrrolic N, graphitic N, and oxidized N. As shown in Figure 5d. The Fe2p spectrum is fitted into six peaks [38,39]: Fe⁰ (710.88 eV), with a relative content of 56.01%; the peaks at 714.36 eV and 727.16 eV correspond to Fe²⁺ 2p₃/₂ and 2p₁/₂, with a relative content of 23.39%; the peaks at 719.04 eV and 731.84 eV correspond to Fe³⁺ 2p₃/₂ and 2p₁/₂, with a relative content of 20.60%; and the peak at 723.68 eV is a satellite peak of Fe²⁺. The iron in Fe-N-C exists in three different forms: Fe⁰, Fe²⁺, and Fe³⁺.

2.3. Performance Evaluation of Fe-N-C Activated PMS

The effect of various pH on the degradation of phenol by the Fe-N-C/PMS system is shown in Figure 6a,b. As the initial pH increases, the phenol degradation rate remains almost the same within 60 min, with the reaction rate constants being 0.1210 min⁻¹, 0.1491 min⁻¹, 0.1054 min⁻¹, and 0.1390 min⁻¹, respectively. Both acidic and alkaline environments slightly inhibit the phenol degradation rate. Like the N-C/PMS system, the phenol degradation rate in the Fe-N-C/PMS system is not significantly affected by the pH of the solution. Most metal-doped catalysts that activate PMS for organic pollutants degradation are typically inhibited in alkaline conditions due to the surface passivation of the doped metal, which reduces active sites. However, this phenomenon was not observed in the Fe-N-C/PMS system, indicating that Fe-N-C material is more adaptable to a wider pH range.

The effects of inorganic anions and natural organic substances on phenol degradation by the Fe-N-C/PMS system are shown in Figure 6c,d. The addition of Cl⁻, SO₄²⁻, and HA all inhibited phenol degradation (60 min) slightly, but the anions reduced the degradation rate. After adding Cl⁻, the system’s reaction rate constantly decreased from 0.1491 min⁻¹ to 0.0804 min⁻¹. This could be due to Cl⁻ reacting with SO₄^•− and •OH to form other low-activity radicals, which in turn affected the phenol degradation rate. After adding SO₄²⁻, the system’s reaction rate constant became 0.1149 min⁻¹, with the inhibition effect being less pronounced. When HCO₃⁻ was added, the system’s reaction rate constant dropped to 0.0422 min⁻¹, and HCO₃⁻ had the most significant inhibitory effect on phenol degradation, reducing the phenol degradation rate to 92%. This may be because HCO₃⁻ competes with phenol for SO₄^•− and •OH, producing HCO₃^•−, which has a lower redox potential. Additionally, when HCO₃⁻ dissociates in solution, it generates CO₃²⁻, which, as an electron donor, also consumes SO₄^•− and •OH, thereby interfering with phenol degradation [40]. After adding HA, the system’s reaction rate constant decreased from 0.1491 min⁻¹ to 0.0774 min⁻¹. It is speculated that HA contains electron-rich groups such as -OH and -COOH, which readily react with active species in the system, competing with phenol for oxidation. Moreover, HA may also be adsorbed on the surface of Fe-N-C, leading to a reduction in the number of active sites on the catalyst surface, thus affecting the adsorption of PMS and phenol [41]. Compared to the N-C/PMS system, the Fe-N-C/PMS system has slightly lower resistance to interference. It is speculated that the difference between the two systems arises from the distinct reaction mechanisms by which the materials activate PMS to degrade organic pollutants.

The stability results of Fe-N-C are shown in Figure 6e. The phenol degradation rates for Fe-N-C in the first and second cycles were 99.71% and 99.54%, respectively. However, in the third cycle, the phenol degradation rate dropped to 80.00%, and it further decreased to 69.73% in the fourth cycle. Under the same condition, the phenol degradation rate for N-C in the third cycle was 45.25%, indicating that compared to N-C, Fe-N-C shows significantly improved reusability and stability. The strategy of doping an appropriate amount of metal elements into the material to enhance its structural stability and reusability is effective. However, the stability of Fe-N-C is still relatively weak compared to other metal materials [42,43,44], and further improvements are needed. The decrease in material activity could be due to the active sites on the material’s surface being covered by intermediate products, pore blockage, or the loss of active components such as iron and nitrogen in the material.

2.4. Mechanism Study of Fe-N-C Activated PMS

The quenching experiment results are shown in Figure 7a,b. Without the addition of quenching agents, the reaction rate constant for phenol degradation in the Fe-N-C/PMS system was 0.1491 min⁻¹. When tert-butyl alcohol, with a molar ratio of 100:1 to the oxidant, was added to selectively scavenge •OH, the system’s reaction rate constant decreased to 0.1174 min⁻¹. When methanol, at the same concentration, was added to quench both •OH and SO₄^•−, the system’s reaction rate constant decreased further to 0.0627 min⁻¹, indicating a stronger inhibitory effect than tert-butyl alcohol. The addition of both quenching agents suppressed the degradation of phenol, suggesting the presence of •OH and SO₄^•− in the reaction system. However, the phenol degradation rate did not decrease, indicating that other active species were also present in the system. When benzoquinone and furfural were added to the solution, the phenol degradation rates after 60 min decreased to 86.32% and 53.46%, indicating the presence of O₂^•− and ¹O₂ in the Fe-N-C/PMS system.

The ESR test results are shown in Figure 7c–e. When only PMS was added, characteristic signal peaks of DMPO-OH and -SO₄^•− were not observed in the system. But, when PMS and Fe-N-C were simultaneously added, characteristic signal peaks for •OH and SO₄^•− were detected. As the reaction time was extended to 10 min, the intensities of the characteristic signal peaks for •OH and SO₄^•− increased, indicating the presence of •OH and SO₄^•− in the solution, and the accumulation of active radicals over time. As shown in Figure 5d, DMPO-O₂^•− characteristic signal peak with a 1:1:1:1:1:1 intensity ratio did not appear when only PMS was added. After the addition of Fe-N-C, a clear DMPO-O₂^•− characteristic signal peak was detected in the system, indicating the presence of O₂^•− in the system. As shown in Figure 7e, when only PMS was added, a characteristic signal peak of TEMP-¹O₂ with an intensity ratio of 1:1:1 was observed, which was attributed to the self-decomposition of PMS [45]. When both PMS and Fe-N-C were simultaneously added to the solution, the intensity of the TEMP-¹O₂ characteristic signal peak was significantly enhanced, indicating that Fe-N-C can activate PMS to generate more ¹O₂.

The above free radical quenching experiments and ESR test confirm that both radical pathways (•OH, SO₄^•−, O₂^•−) and non-radical pathways (¹O₂) coexist in the Fe-N-C/PMS system, with •OH, SO₄^•−, and ¹O₂ playing dominant roles in the system. However, the degradation mechanism of PMS activation in the Fe-N-C/PMS system remains unclear. Therefore, XPS comparative analysis was conducted on the Fe-N-C after the reaction (Figure 8a), where the relative element contents of C, O, N, and Fe are 83.18%, 13.79%, 1.93%, and 1.10%. Compared to the relative element contents of Fe-N-C before use, the relative contents of N and Fe decreased from 3.31% and 1.26% to 1.93% and 1.10%, respectively. This suggests that N and Fe elements in the material are the main active sites involved in the reaction.

The fitting peaks further analyze the changes in the oxidation states of C, N, and Fe elements. Figure 8b and

Table 2. Morphological changes of C before and after Fe-N-C reaction.

Morphological Changes of C	C=C	C-C	C-O/C-N	C=O
Before	36.92%	43.62%	11.48%	7.98%
After	35.99%	48.78%	9.63%	5.60%

show the high-resolution C1s spectrum. After the reaction, the relative content of C=C in Fe-N-C decreased, while the relative content of C-C increased. Both sp²-hybridized carbon and sp³-hybridized carbon can weaken the O-O bond in PMS, promoting the generation of •OH or SO₄^•−. However, due to the imbalance of electrons, sp²-hybridized carbon is more active than sp³-hybridized carbon [46]. The relative content of C-O/C-N and C=O decreased from 11.48% and 7.98% to 9.63% and 5.60%, respectively, indicating that both act as active sites in the reaction. The C=O, due to its lone pair of electrons, can enhance the electron density of carbon near the Fe-N-C/PMS interface [47]. Figure 8c and

Table 3. Morphological changes of N before and after Fe-N-C reaction.

Morphological Changes of N	Pyridinic N	Pyrrolic N	Graphitic N	Oxidized N
Before	15.22%	28.10%	40.53%	16.15%
After	21.46%	52.10%	16.25%	10.20%

present the high-resolution N1s spectrum. After the reaction, the relative content of graphite nitrogen decreased from 40.53% to 16.25%, suggesting that graphite nitrogen can serve as an active site in the reaction. Graphitic nitrogen can disrupt the electron density of sp²-hybridized carbon, making adjacent carbon atoms positively charged and weakening the O-O single bond of PMS adsorbed at this site. This causes PMS to enter a metastable state, leading to electron rearrangement and the generation of •OH or SO₄^•−. Figure 8d and

Table 4. Morphological changes of Fe before and after Fe-N-C reaction.

Morphological Changes of Fe	Fe0	Fe2+	Fe3+
Before	56.01%	23.39%	20.60%
After	47.81%	27.25%	24.95%

show the high-resolution Fe2p spectrum. After the reaction, the relative content of Fe⁰ on the surface of Fe-N-C decreased from 56.01% to 47.81%, indicating that Fe⁰ is an important active site in the reaction, and part of Fe⁰ is converted to Fe²⁺ and Fe³⁺. The relative contents of Fe²⁺ and Fe³⁺ increased from 23.39% and 20.60% to 27.25% and 24.95%, respectively, maintaining a relatively balanced ratio, suggesting the presence of an Fe²⁺/Fe³⁺ cycling process.

At the same time, the decrease in catalyst activity after the cyclic experiments can be explained by XPS analysis. After each cycle, the catalyst undergoes slight oxidation, and the N and Fe elements, which are the main active sites, gradually decrease. This may lead to reduced catalytic activity after multiple cycles. Additionally, during the catalyst recovery, it is possible that intermediate products covering the catalytic active sites were not fully removed, leading to a reduction in catalytic activity. Finally, multiple cycles may cause structural changes in the catalyst, such as a decrease in active sites or changes in pore size distribution, which can result in decreased catalytic activity.

Based on the above discussion, a possible reaction mechanism for the degradation of phenol by PMS activated by Fe-N-C is proposed. (i) PMS decomposes itself to generate ¹O₂; (ii) The cycle of Fe⁰ and Fe²⁺/Fe³⁺ in Fe-N-C activates PMS to produce SO₄^•− and •OH, while dissolved oxygen gains electrons to generate O₂^•− [48,49]; then, with the participation of water molecules, PMS is activated by sp2-hybridized carbon, C-O/C-N, C=O, and graphite nitrogen to produce SO₄^•−, •OH, O₂^•−, and ¹O₂; (iii) O₂^•− further reacts with •OH to generate ¹O₂ [49], which can explain the ESR experimental results. In conclusion, the combination of free radical pathway dominated by SO₄^•−, •OH, and O₂^•−, with non-radical pathway dominated by ¹O₂, jointly promotes the degradation of phenol. The relevant reaction equations are shown below (Equations (1)–(9)).

$${\mathrm{SO}}_5^{2-}+{\mathrm{HSO}}_5^{-} \to {}^1\mathrm{O}_2{+{\mathrm{HSO}}_4^{-}-\mathrm{SO}}_4^{2-}$$

(1)

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

(2)

$${\mathrm{Fe}}^0+2{\mathrm{HSO}}_5^{-} \to {\mathrm{Fe}}^{2+}+{ 2\mathrm{SO}}_4^{•-}+2\mathrm{O}{\mathrm{H}}^{-}$$

(3)

$${\mathrm{Fe}}^{2+}+{ \mathrm{HSO}}_5^{-} \to {\mathrm{Fe}}^{3+}+{ \mathrm{SO}}_4^{•-}+\mathrm{O}{\mathrm{H}}^{-}$$

(4)

$${\mathrm{Fe}}^{3+}+{ \mathrm{HSO}}_5^{-} \to { \mathrm{Fe}}^{2+}+{ \mathrm{SO}}_5^{•-}+{\mathrm{H}}^{+}$$

(5)

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

(6)

$${\mathrm{Fe}}^{2+}+{ \mathrm{O}}_2 \to {\mathrm{Fe}}^{3+}+{ \mathrm{O}}_2^{•-}$$

(7)

$${\mathrm{e}}^{-}+ {\mathrm{O}}_2 \to { \mathrm{O}}_2^{•-}$$

(8)

$${\mathrm{O}}_2^{•-}+•\mathrm{OH} \to {}^1\mathrm{O}_2+\mathrm{O}{\mathrm{H}}^{-}$$

(9)

2.5. Material Comparison

To better evaluate the performance of this material for the degradation of organic pollutants, a comparison was made between Fe-N-C and metal materials from the literature in terms of their effectiveness in activating PMS for phenol degradation (as shown in

Table 5. Comparison of phenol removal performance of different materials.

Materials	Reaction Conditions	Degradation Effect	Reference
Fe3C-N-CNT	[phenol] = 20 mg/L; [catalyst] = 200 mg/L; [PMS] = 3 mM	100% 20 min	[24]
γ-MnO2-ZnFe2O4-rGO	[phenol] = 20 mg/L; [catalyst] = 200 mg/L; [PMS] = 3 mM	100% 30 min	[50]
Fe0-Fe3C-CS	[phenol] = 10 mg/L; [catalyst] = 20 mg/L; [PMS] = 3 mM	100% 10 min	[51]
Fe-MOF	[phenol] = 20 mg/L; [catalyst] = 200 mg/L; [PMS] = 3 mM	100% 60 min	[52]
Fe-N-C	[phenol] = 50 mg/L; [catalyst] = 150 mg/L; [PMS] = 1.5 mM	100% 60 min	This thesis

).

3. Materials and Methods

3.1. Chemical Reagents

Humic acid (HA), sodium bicarbonate (NaHCO₃), potassium chloride (KCl), urea, furfural alcohol (FFA), p-benzoquinone (p-BQ), ammonia solution (NH₃·H₂O), 2,2′-azino-bis(3-ethylbenzothiazoline-6-sulfonic acid) diammonium salt (ABTS) were purchased from Shanghai Macklin Biochemical Technology Co., Ltd.(Shanghai, China); phenol, sodium hydroxide (NaOH), hydrochloric acid (HCl), sulfuric acid (H₂SO₄) were purchased from Sinopharm Chemical Reagent Co., Ltd. (Shanghai, China); ferric chloride hexahydrate (FeCl₃·6H₂O), potassium hydrogen persulfate (KHSO₅), sodium persulfate (Na₂S₂O₈), hydrogen peroxide (H₂O₂), anhydrous sodium sulfate (Na₂SO₄), tert-butyl alcohol (TBA) were purchased from Shanghai Aladdin Biochemical Technology Co., Ltd. (Shanghai, China); methanol (MeOH) was purchased from TEDIA (Fairfield, OH, USA).

3.2. Material Preparation

A certain amount of urea was heated at a rate of 5 °C/min to 550 °C and kept for 4 h to obtain a light-yellow sample, which is graphite-like carbon nitride (g-C₃N₄).

For the preparation of nitrogen-doped carbon materials, HA, g-C₃N₄, and FeCl₃·6H₂O were used as raw materials. The synthesis method is as follows: HA, g-C₃N₄, and FeCl₃·6H₂O were mixed in a mass ratio of 1: 2: 0.8 and placed in 40 mL of deionized water with magnetic stirring. Afterward, the solution was evaporated to dryness, and the resulting mixture was placed in a tube furnace. Under a nitrogen atmosphere, the temperature was raised to 300 °C and held for 1 h, then further increased to 800 °C and held for 1 h. After the sample cooled, the product was washed with 1 mol/L hydrochloric acid, followed by water washing until neutral. The material was dried and ground, resulting in the Fe-N-C material. The nitrogen-doped carbon material prepared from humic acid and g-C₃N₄ as raw materials is denoted as N-C.

3.3. Material Characterization

The material’s crystal structure was tested using a SmartLab SE X-ray diffractometer (XRD, Rigaku, Tokyo, Japan). The nitrogen adsorption–desorption isotherm of the material was measured using a physical adsorption analyzer (Quantachrome Instruments, Boynton Beach, FL, USA). The surface morphology of the material was characterized by scanning electron microscopy (SEM) and energy-dispersive spectroscopy (EDS) using a Regulus 8100 microscope (Hitachi, Tokyo Japan). The surface elemental composition and chemical states of the material were analyzed with a K-Alpha X-ray photoelectron spectrometer (XPS, Thermo Scientific, Waltham, MA, USA). The synthesized material was characterized by various techniques, and the operating parameters are listed in Text S1. The degree of graphitization and defect levels of the material were examined using a inVia Raman spectrometer (Renishaw, London, UK). The elemental composition of the sample was analyzed by an elemental analyzer (Elementar, Rhine Main, Germany). Pollutant concentrations were determined using high-performance liquid chromatography (HPLC), and the concentration of PMS was measured by the Co²⁺/PMS-ABTS colorimetric method. The active species in the system were measured by ESR spectroscopy using a EMXplus ESR spectrometer (Bruker, Saarbrücken, Germany). DMPO was used as an electron spin trap agent to detect •OH and SO₄^•− in aqueous solution, and O₂•⁻ in methanol solution. TEMP was used as an electron spin trap agent to detect ¹O₂ in aqueous solution.

3.4. Material Performance Evaluation

A phenol solution with the specified concentration was prepared, and a measured amount of PMS and catalyst were sequentially added to initiate the activation reaction. The mixture was placed on a magnetic stirrer for stirring. At specified time intervals, 2 mL samples were taken using a syringe and filtered through a 0.45 μm micropore filter for pollutant concentration analysis. After the reaction, the solid was collected and dried in an oven. The above procedures were repeated four times to complete the material’s reusability test. NaOH and H₂SO₄ solutions were employed to adjust the pH of the solution, and no buffer solution was added during the process.

4. Conclusions

In this study, we report an iron and nitrogen doped HA-based carbon catalyst (Fe-N-C) synthesized from HA, g-C₃N₄, and FeCl₃·6H₂O. Fe-N-C/PMS system exhibits excellent performance on phenol degradation across a wide pH range and in complex water treatment environments. The characterization research results show that iron doping effectively alters the morphology and size of Fe-N-C. At the same time, iron-nitrogen co-doping provides abundant active N and Fe sites in Fe-N-C. Fe exists in the form of Fe⁰ and Fe₃C in Fe-N-C, and during the reaction, Fe⁰, Fe²⁺/Fe³⁺ are the main active sites that promote PMS activation. Furthermore, quenching experiments and ESR confirm the presence of singlet oxygen in the Fe-N-C/PMS system. Also discussed are the potential active sites for activating PMS and the reaction mechanisms. This study provides a promising low-cost iron-nitrogen co-doped HA-based carbon catalyst for water environmental remediation and the degradation of organic pollutants in wastewater.