New Insights About the Role of Some Activated Carbon Materials in the H 2 O 2 -Involved Advanced Oxidation of Organic Contaminants

Abstract

In this study, we investigated the degradation of tetracycline (TC), rhodamine B (RhB), and p-nitrophenol (PNP) in the activated carbon (AC)/H₂O₂ system using three AC materials prepared from coconut shell, glucose, and corn straw. Our findings indicate that, despite the significant decomposition of H₂O₂, the degradation of TC and RhB in the AC/H₂O₂ system was almost negligible. For PNP, a weak degradation was observed in the AC/H₂O₂ system, but the degradation rate was much slower than that obtained from the homogeneous Fenton oxidation (H₂O₂/Fe²⁺) system under the same amount of H₂O₂ consumption. Only weak electron spin resonance signals of DMPO-·OH and DMPO-·O₂⁻ were detected from the AC/H₂O₂ system. This suggested that H₂O₂ decomposition mainly followed a nonradical pathway in the system, consistent with the poor performance of organic degradation in spite of significant H₂O₂ decomposition. In summary, this study demonstrated that some AC materials may be not good Fenton-like catalysts, and while AC is sometime used as heterogeneous support of Fenton-like catalysts, the role of AC in the H₂O₂−invovled advanced oxidation should be critically examined, as it may not facilitate H₂O₂ utilization geared to organic contaminant degradation.

Introduction

Activated carbon (AC) is always considered as a good adsorbent widely used in eliminating various pollutants in the environmental protection field. However, its new application as a catalyst or catalyst carrier used in H₂O₂−involved advanced oxidation processes (AOPs) for the degradation of organic contaminants attracted wide attention over the past 15 years (Anfruns, et al., 2014; Duarte et al., 2013; Li et al., 2022; Li et al., 2023; Santos et al., 2009; Zarate-Guzmán et al., 2019). Among various applications of AC in H₂O₂−involved AOPs, the typical one is to develop various heterogeneous Fenton-like processes (Anfruns, et al., 2014; Duarte et al., 2013; Li et al., 2022; Li et al., 2023; Santos et al., 2009; Zarate-Guzmán et al., 2019).

It is well known that the traditional Fenton process based on the homogeneous reaction between H₂O₂ and Fe²⁺ generating hydroxyl radical (·OH) has been successfully applied in the treatment of organic industrial wastewater (Pignatello et al., 2006). However, the homogeneous Fenton process suffers from several disadvantages including a narrow pH application range (pH <3), high consumption of chemicals, and production of iron-containing sludge (Bokare and Choi, 2014; Luo et al., 2021; Pignatello et al., 2006). To overcome these disadvantages, the heterogeneous Fenton-like processes characterized by the application of solid catalysts are attracting more and more attention (Bokare and Choi, 2014; Luo et al., 2021). Usually, the solid catalysts used in heterogeneous Fenton-like processes mainly include various iron oxides such as magnetite (Fe₃O₄) (He et al., 2014), hematite (α-Fe₂O₃) (Wang et al., 2024), goethite (α-FeOOH) (Wu et al., 2021), and iron oxychloride (FeOCl) (Zhang et al., 2020), which are responsible for H₂O₂ activation. However, numerous studies also revealed that AC can be utilized in the heterogeneous Fenton-like processes as a catalyst to activate H₂O₂ generating ·OH for organics degradation (Anfruns, et al., 2014; Li et al., 2023; Santos et al., 2009) or as a catalyst carrier to load and disperse catalyst particles preventing coagulation for attaining high catalytic activity (Duarte et al., 2013; Li et al., 2022; Zarate-Guzmán et al., 2019).

In fact, it has long been recognized that AC can give rise to H₂O₂ decomposition as a catalyst. Although the reaction mechanism of H₂O₂ decomposition by AC is not fully understood, its application in degrading various organic pollutants such as dyes (Santos et al., 2009), bisphenol A (Li et al., 2023), pentachlorobenzene (Takaoka et al., 2007), and 4-chlorophenol (Huang et al., 2003) has been reported. These studies confirmed that H₂O₂ decomposition over AC surface can generate ·OH resulting in organics degradation. AC surface is considered as an electron-transfer mediator to promote ·OH formation, and its catalytic behavior is similar to Fe(II) species in the Fenton reagent (Anfruns et al., 2014). Besides, more extensive application of AC in H₂O_2⁻involved AOPs was to prepare heterogeneous Fenton catalysts as a catalyst carrier due to its huge specific surface area, high stability, and variable surface chemical property. These catalysts such as α-FeOOH/AC (Li, et al., 2022) and Fe₃O₄/AC (Jaafarzadeh et al., 2015) exhibited considerable catalytic activity to degrade organic contaminants through activating H₂O₂. The utilization of AC can not only prevent the coagulation of catalyst particles but also accelerate the conversion of Fe(III) to Fe(II) in iron oxides enhancing H₂O₂ activation attributed to some functional groups such as hydroxyl groups and quinone groups (Li et al., 2022; Yan et al., 2022).

Besides organics degradation performance, H₂O₂ utilization efficiency is another important issue concerning technical and economic feasibility for the practical application of those H₂O_2⁻involved AOPs. Although many studies reported the positive effects of AC in the H₂O_2⁻involved AOPs, until now, there has been no study focused on the H₂O₂ utilization efficiency of these processes (Anfruns, et al., 2014; AZarate-Guzmán et al., 2019; Duarte et al., 2013; Li et al., 2022; Li et al., 2023; He et al., 2022; Molamahmood et al., 2022; Rey et al., 2016; Santos et al., 2009). Thus, in this work, the three AC materials prepared, respectively, with coconut shell, glucose, and corn straw were utilized to activate H₂O₂ for the degradation of tetracycline (TC), rhodamine B (RhB), and p-nitrophenol (PNP). The degradation results were compared with those obtained from conventional Fenton oxidation process. Moreover, the comparison of H₂O₂ utilization efficiency was emphasized to objectively evaluate the real significance of utilizing AC in the H₂O_2⁻involved AOPs.

Experimental Section

Materials

All chemicals were of analytical grade and used without further purification. All solutions were prepared using deionized water with conductivity <0.4 μS/cm. Aqueous solutions of H₂SO₄ and NaOH were adopted to adjust pH values. One commercialized AC product (1# sample) was purchased from Hainan Xinghuo Activated Carbon Co., Ltd. (China).

Preparation and characterization of AC materials

Glucose and corn straw were used to prepare the two AC materials, and the obtained products were labeled as 2# and 3# samples, respectively. The preparation involved the carbonization of glucose and corn straw, followed by the activation treatment. The carbonization of glucose was conducted by one hydrothermal process. Specifically, 12.0 g glucose was dissolved into 70 mL deionized water, and then this solution was put into a 100 mL PTFE-lined reactor for the hydrothermal treatment at 180°C for 12 h. After cooling, the formed carbon material was centrifugally washed using ethanol and deionized water successively. Afterward, the carbon material was dried at 105°C for 2 h in an air circulation oven for the subsequent activation treatment. For the carbonization of corn straw, firstly, it was washed with deionized water and dried at 105°C and then was ground to powder. After that, corn straw powder was subjected to hydrothermal treatment in a muffle furnace at 550°C for 2 h with a heating rate of 10°C/min under nitrogen protection to synthesize another carbon material. The activation treatment of the above two carbon materials was conducted by the same procedures. Firstly, 0.8 g carbon and 3.2 g KOH were well mixed and then subjected to hydrothermal treatment in a muffle furnace at 600°C for 2 h with a heating rate of 20°C/min under nitrogen protection. After that, the above mixture was washed with deionized water till the pH value of the washing liquor was close to 7.0. Finally, the obtained AC was dried at 105°C for 2 h.

Scanning electron microscopy (SEM) images were obtained with a JSM-6700F electron microscope. The crystal structure of AC samples was characterized by X-ray diffraction (XRD) with a Rikagu D/MAX-2500 diffractometer. The functional groups of AC samples were analyzed by Fourier transform infrared (FTIR) spectroscopy using a Bruker FTIR-Vertex 70 analyzer. The specific surface area was measured via the N₂ adsorption–desorption method on a Micromeritics ASAP2020 instrument (USA).

Experiments of H₂O₂ decomposition

H₂O₂ decomposition by the three AC samples was investigated at different pH values. Typically, 100 mL H₂O₂ solution (100 mg/L) was transferred into a glass reactor, followed by pH adjustment, and then a certain amount of AC was added into the solution to start the catalytic decomposition of H₂O₂ under magnetic stirring. At specified time intervals, 2 mL of the suspension was extracted and filtered through a 0.22 μm membrane for H₂O₂ determination. The H₂O₂ decomposition efficiency was calculated by the equation: [(C₀ − C)/C₀] × 100%, where C₀ and C were H₂O₂ concentration values before and after the reaction. Because the three ACs have significant difference in H₂O₂ decomposition capability as the same dosage adopted, the H₂O₂ decomposition experiments were performed with 3.0–10.0 g/L 1# AC, 0.6 g/L 2# AC, and 1.0 g/L 3# AC to obtain appropriate H₂O₂ decay, facilitating the study of the relationship between H₂O₂ decomposition and organics degradation.

Degradation experiments

The degradation of TC by the three AC materials (1#, 2#, and 3#) was carried out at pH = 3.0, 7.0, and 9.0, respectively. The dosages of 1# AC, 2# AC, and 3# AC used were 3.0, 0.6, and 1.0 g/L, respectively. For each degradation experiment, 100 mL TC solution (60 mg/L) was transferred into a glass reactor, followed by pH adjustment, and then AC was added into the solution to perform the adsorption for 60 min under magnetic stirring. After that, 10 mg H₂O₂ was added into the solution to start the degradation with the initial H₂O₂ concentration of 100 mg/L. At specified time intervals, samples were drawn and filtered through a 0.22 μm membrane for H₂O₂ and TC determination. Furthermore, the degradation of 40 mg/L RhB by 0.6 g/L 2# AC and the degradation of 100 mg/L PNP by 0.4 g/L 2# AC were conducted according to the same experimental procedures. 2# AC exhibited very strong ability toward PNP adsorption, and thus a lower 2# AC dosage of 0.4 g/L was used for PNP degradation to facilitate degradation performance analysis.

For the purpose of comparison, the degradation of TC, RhB, and PNP by the conventional Fenton oxidation system (H₂O₂/Fe²⁺) was carried out at pH = 3.0 with a fixed molar ratio of H₂O₂ to Fe²⁺ (about 5:1). The TC removal efficiency was calculated by the equation: [(C₀ − C)/C₀] × 100%, where C₀ and C were TC concentration values before and after the reaction.

Electron spin resonance tests

The generation of ·OH and ·O_2⁻ in H₂O₂/AC system and H₂O₂/Fe²⁺ system was comparatively studied by electron spin resonance (ESR) tests (Qu et al., 2022; Tian et al., 2022). To identify ·OH generated from H₂O₂/AC system, 5 mg 2# AC sample, 0.5 mg H₂O₂, and 10 μL DMPO were added into 5 mL deionized water (pH = 7.0), and then 50 µL sample was used to record DMPO-·OH ESR signal at 4, 8, and 12 min on an electron paramagnetic resonance spectrometer (Bruker ESR A300). To identify ·OH generated from H₂O₂/Fe²⁺ system, 0.8 mg FeSO₄·7H₂O, 0.5 mg H₂O₂, and 10 μL DMPO were added into 5 mL deionized water (pH = 3.0) to prepare the sample for the test. To record DMPO-·O₂⁻ ESR signals, similar tests were carried out with methanol as the dispersion phase instead of water.

Analytical methods

A PHS-25C pH meter was used to measure solution pH values. H₂O₂ concentration was measured by the spectrophotometric method according to the light absorption of titanium–hydrogen peroxide complex at 410 nm. TC and RhB concentrations were determined by the spectrophotometric method at 357 and 554 nm (Zhang et al., 2019; Wang et al., 2022), respectively. PNP concentration was determined by gas chromatography and the detailed analysis conditions were provided in the reference (Chu et al., 2020).

Results and Discussion

Characterization of various AC materials

The three AC materials were used in this study including one commercialized product produced with coconut shell and two self-made products prepared with glucose and corn straw. The commercialized AC was columnar with a diameter of about 2 mm and a height of about 3–8 mm, and the other two AC were powder materials (Fig. 1a). The SEM image of the commercialized AC (1#) exhibited its dense, smooth, and porous structure (Fig. 1b). The AC prepared with glucose (2#) possessed a loose and porous structure (Fig. 1c). The AC prepared with corn straw (3#) possessed an unordered and porous structure and still remained an identifiable texture structure (Fig. 1d).

FIG. 1.

Photographs of the three AC materials (a). SEM images of 1# AC sample (b), 2# AC sample (c), and 3# AC sample (d). XRD patterns of the three AC materials in the 2θ range of 10°–80° (e) and FTIR patterns in the wavenumber range of 500–3875/cm (f). AC, activated carbon; SEM, scanning electron microscopy; XRD, X-ray diffraction.

Figure 1e shows the XRD spectra of the three AC materials. The two broad diffraction peaks can be observed over the 2θ range of 20–30° and 40–50° for the three materials, which are typical peaks of disordered carbon structure, indicating the amorphous nature of the materials (Li et al., 2022; Wazir et al., 2023; Yao et al., 2023), but with a microcrystalline structure similar to that of graphite within a certain range (Alhnidi et al., 2020; Yao et al., 2023). For adsorption purposes, an amorphous structure can facilitate the entrance of adsorbate molecules because of more empty spaces. The XRD spectrum of 2# AC displays the two narrow impurity diffraction peaks at 28.3° and 34.5°, which may be related to a small amount of residual glucose. Figure 1f shows that the three materials have the same FTIR absorption peaks. The characteristic peak at 3436 cm⁻¹ is ascribed to the O–H vibration, while the two peaks at 2925 and 2851 cm⁻¹ are ascribed to the C–H vibration stretching vibration in the aliphatic groups (Li et al., 2011; Li et al., 2022). The peak at 1631 cm⁻¹ is ascribed to the stretching vibration of C=O in carbonyl groups (Liu et al., 2020). The two peaks at 1383 and 1040 cm⁻¹ correspond to the stretching vibration of C–O (Liu et al., 2023; Niazi et al., 2018). These absorption peaks indicate that the three materials are abundant in various functional groups such as hydroxyl, aliphatic, lactone, aldehyde, and carboxyl groups (Niazi et al., 2018). These functional groups may evidently affect the adsorption ability and catalytic activity of AC materials (Anfruns et al., 2014; Vega and Valdés, 2018; Wazir et al., 2023).

H₂O₂ decomposition caused by various AC materials

As the three AC materials dipped into a high-concentration H₂O₂ solution, a large number of oxygen bubbles generated violently at the surface of AC materials, which visually displayed the catalytic decomposition of H₂O₂ caused by various AC materials. To investigate the H₂O₂ decomposition quantitatively, the decomposition of 100 mg/L H₂O₂ was conducted with 3.0–10.0 g/L 1# AC, 0.6 g/L 2# AC, and 1.0 g/L 3# AC. Figure 2a shows the influence of 1# AC dosage on H₂O₂ decomposition at pH = 7.0. Evidently, the higher 1# AC dosage resulted in the faster H₂O₂ decomposition, and a H₂O₂ decomposition efficiency of over 80% was obtained at 150 min as 10.0 g/L 1# AC used. Figure 2b shows the results of H₂O₂ decomposition obtained with 10.0 g/L 1# AC at different solution pH values, indicating that the increase of pH caused the H₂O₂ decomposition speed increased slightly, which was also confirmed by the results of H₂O₂ decomposition obtained with 2# (Fig. 2c) and 3# AC materials (Fig. 2d). In fact, compared with 1# AC, 2# AC and 3# AC were able to cause much faster H₂O₂ decomposition if the same AC dosage was adopted. Specifically, H₂O₂ decomposition efficiencies at the reaction time of 60 min obtained with 10.0 g/L 2# AC and 3# AC at pH = 3.0 reached 100% and 82.5%, respectively. Therefore, the H₂O₂ decomposition capability of the three ACs increased in the order: 1#<3#<2#. The H₂O₂ decomposition performance of 2# and 3# AC was further investigated under the condition of lower dosages: 0.6 g/L 2# AC and 1.0 g/L 3# AC. Further comparison confirmed that 2# AC possessed the fastest H₂O₂ decomposition performance (Fig. 2c and d). The BET surface areas of 1#, 2#, and 3# AC materials were 824, 1025, and 487 m²/g, respectively. Usually, the higher surface area can present more active reactive sites facilitating H₂O₂ decomposition. Besides, three other factors can obviously affect H₂O₂ decomposition such as surface chemical groups of AC materials, metal impurities, and mass transfer efficiency (Anfruns et al., 2014; Georgi and Kopinke, 2005; Vega and Valdés, 2018). The relatively lower H₂O₂ decomposition performance of 1# AC may be mainly ascribed to the low mass transfer efficiency caused by its columnar granular state instead of powder state.

FIG. 2.

Results of H₂O₂ decomposition obtained with different amount of 1# AC at pH = 7.0 (a). Influence of pH on the H₂O₂ decomposition by 1# AC of 10.0 g/L (b), 2# AC of 0.6 g/L (c), and 3# AC of 1.0 g/L (d).

TC, RhB, and PNP degradation by AC/H₂O₂ system and Fenton oxidation system

To comprehensively learn the performance of the AC/H₂O₂ system for organics degradation, the three typical organic compounds including TC, RhB, and PNP were selected for the degradation experiments with the above-mentioned three AC materials. Figure 3a–c shows the concentration changes of TC and H₂O₂ in 1# AC/H₂O₂ system obtained at pH = 3.0, 7.0, and 9.0, respectively. According to Figure 3a–c, the adsorption stage of 60 min caused a noticeable decrease in TC concentration, and the adsorption equilibrium was reached approximatively prior to H₂O₂ addition. However, TC concentration only decreased slightly after H₂O₂ addition in the degradation stage of 180 min, even though the H₂O₂ decomposition efficiency reached 33.5–44.0%. It was evident that substantial H₂O₂ decomposition caused by 1# AC was unable to present effective TC degradation, and this conclusion could also be drawn based on the concentration changes of TC and H₂O₂ in 2# and 3# AC/H₂O₂ systems (Fig. 3d–i).

FIG. 3.

Degradation of TC by the AC/H₂O₂ system with the three AC materials at different pH and the corresponding H₂O₂ decay: 1# AC at pH = 3.0 (a), 7.0 (b), and 9.0 (c); 2# AC at pH = 3.0 (d), 7.0 (e), and 9.0 (f); 3# AC at pH = 3.0 (g), 7.0 (h), and 9.0 (i). TC, tetracycline.

In the above experiments, the lowest and the highest H₂O₂ decomposition efficiencies were 19.0% (Fig. 3g) and 44.0% (Fig. 3b), respectively, which denoted the consumed H₂O₂ concentration in all the degradation experiments ranged from 19.0 to 44.0 mg/L. Such a H₂O₂ consumption in the AC/H₂O₂ system presented nearly negligible TC degradation. Thus, an important question of whether such a H₂O₂ consumption can present effective TC degradation in other H₂O_2⁻involved AOPs was naturally raised. To answer this question, the degradation of 60 mg/L TC by the conventional Fenton system (H₂O₂/Fe²⁺) was conducted at pH = 3.0 with H₂O₂ dosages of 10–60 mg/L. Figure 4a–d shows the concentration changes of TC and H₂O₂ obtained from the H₂O₂/Fe²⁺ system. It was found that the increase in H₂O₂ dosage greatly enhanced TC degradation speed. As 10, 20, 40, and 60 mg/L H₂O₂ was adopted for the degradation, the Fenton reactions consumed 4.2, 12.3, 35.2, and 54.5 mg/L H₂O₂ within 120 min, respectively, and the corresponding TC removal efficiencies of 46.1%, 74.0%, 88.9%, and 94.9% were achieved, indicating effective TC degradation. On comparing these results with those obtained from the AC/H₂O₂ system, it could be concluded that the H₂O₂ utilization efficiency for TC degradation in the AC/H₂O₂ system was much lower than that in the H₂O₂/Fe²⁺ system. In other words, a large amount of H₂O₂ was consumed uselessly in the AC/H₂O₂ system considering TC degradation performance.

FIG. 4.

Degradation of TC by the H₂O₂/Fe²⁺ system with different dosages of H₂O₂ and Fe²⁺: 10 mg/L H₂O₂ and 3.3 mg/L Fe²⁺ (a); 20 mg/L H₂O₂ and 6.6 mg/L Fe²⁺ (b); 40 mg/L H₂O₂ and 13.2 mg/L Fe²⁺ (c); and 60 mg/L H₂O₂ and 19.8 mg/L Fe²⁺ (d).

To further learn the performance of the AC/H₂O₂ system toward organics degradation, 2# AC was used in this oxidation system for the degradation of 40 mg/L RhB and 100 mg/L PNP. As displayed in Figure 5a, after reaching the adsorption equilibrium, the RhB concentration decreased very slowly exhibiting nearly negligible RhB degradation under different pH conditions, which was similar to the TC degradation by the AC/H₂O₂ system. Additionally, the H₂O₂ concentration decreased by 42.0 mg/L at least after the degradation for 150 min. Figure 5b displays the degradation of RhB by the H₂O₂/Fe²⁺ system at pH = 3.0. It was found that RhB concentration decreased rapidly within 120 min as 40 and 60 mg/L H₂O₂ was used for the degradation. These results suggested that the H₂O₂/Fe²⁺ system presented much faster RhB degradation than the AC/H₂O₂ system under the condition of approximate H₂O₂ consumption.

FIG. 5.

Degradation of RhB by the AC/H₂O₂ system with 0.6 g/L 2# AC and 100 mg/L H₂O₂ at different pH (a). Degradation of RhB by the H₂O₂/Fe²⁺ system with different dosages of H₂O₂ and Fe²⁺ (b). Comparative degradation of PNP by the two systems (c) and the corresponding pseudo-second-order kinetic curves (d).

The results of PNP degradation obtained from the AC/H₂O₂ system and the H₂O₂/Fe²⁺ system at pH = 3.0 are shown in Figure 5c. It was found that 2# AC possessed a powerful capacity toward PNP adsorption, and 2# AC of 0.4 g/L caused the PNP concentration to decrease to about 20 mg/L from the initial concentration of 100 mg/L. Prior to adding H₂O₂, adsorption equilibrium was approximately reached by the adsorption of 60 min. It was found that the degradation of PNP by the AC/H₂O₂ system for 120 min caused the concentration to decrease to 8.7 mg/L from about 20 mg/L and the H₂O₂ concentration decreased to 60 mg/L from the initial concentration of 100 mg/L. Thus, for the purpose of comparison, the degradation of 20 mg/L PNP by the H₂O₂/Fe²⁺ system with 40 mg/L H₂O₂ and 17.8 mg/L Fe²⁺ was conducted. PNP concentration decreased to 3.0 mg/L after degradation for 120 min, accompanied by complete H₂O₂ consumption. The degradation results in Figure 5c were further analyzed using the pseudo-second-order reaction kinetics. As shown in Figure 5d, the second-order rate constant k obtained from the AC/H₂O₂ system was 0.0006 L/mg·min, which was much lower than that obtained from the H₂O₂/Fe²⁺ system (0.0024 L/mg·min). Evidently, although the AC/H₂O₂ system was able to present a weak PNP degradation, its performance toward PNP degradation was far weaker than that of the H₂O₂/Fe²⁺ system. Based on the above results, it can be concluded that these ACs are not effective Fenton-like catalysts for organics degradation. This finding appears to be different from that reported in some studies, in which the AC/H₂O₂ system exhibited a certain capability for organics degradation. Besides the difference in surface chemical groups, the different contents of metal oxides in ACs may be another important reason for the different degradation performance (Anfruns et al., 2014; Fang et al., 2014). Usually, some metal oxides (e.g., CuO, Fe₂O₃) that existed in ACs can be beneficial to organics degradation in the presence of H₂O₂ (Wang and Tang, 2021).

Mechanism discussion

Numerous studies have revealed that the degradation of organic contaminants by AOPs mainly relies on the generation of various reactive oxygen species such as ·OH and superoxide anion radical (·O₂⁻) (Dong et al., 2022; Luo et al., 2021). For example, ·OH is usually considered as the main reactive oxygen in the H₂O₂/Fe²⁺ system generated via the classical Fenton reaction [Eq. (1)]. In this work, the generation of ·OH and ·O₂⁻ by the AC/H₂O₂ system and the H₂O₂/Fe²⁺ system was studied comparatively by ESR tests using the same amount of H₂O₂. As displayed in Figure 6a–d, the ESR signals from both the AC/H₂O₂ system and the H₂O₂/Fe²⁺ system presented the characteristic peaks of DMPO-·OH and DMPO-·O₂⁻ demonstrating the generation of ·OH and ·O₂⁻ in the two systems. However, the intensity of ESR signals from the AC/H₂O₂ system (Fig. 6b and d) was much weaker than that from the H₂O₂/Fe²⁺ system (Fig. 6a and 6c), indicating that the AC/H₂O₂ system has a very limited ability for the generation of ·OH and ·O₂⁻, which can explain why the AC/H₂O₂ system did not present effective organics degradation.

FIG. 6.

ESR signals of DMPO-·OH obtained from the H₂O₂/Fe²⁺ system (a) and the AC/H₂O₂ system (b). ESR signals of DMPO-·O₂^– obtained from the H₂O₂/Fe²⁺ system (c) and the AC/H₂O₂ system (d).

Currently, unlike the H₂O₂ decomposition in the H₂O₂/Fe²⁺ system via the classical Fenton reactions generating ·OH and ·O₂⁻ (Eqs. 1 –3), the reaction mechanism of H₂O₂ decomposition at AC surface to generate active oxygen radicals is not fully understood (Dong et al., 2022; Fang et al., 2014; Luo et al., 2021). In some publications, one reaction mechanism similar to the Fenton mechanism was proposed to explain the H₂O₂ decomposition at AC surface (Rey et al., 2011; Zhou et al., 2019). In this mechanism, ·OH and ·OOH are considered as intermediates, while AC and AC⁺ are considered as reduced and oxidized catalyst states, respectively (Eqs. 4 and 5). Furthermore, some publications reported another reaction mechanism following a nonradical pathway, in which H₂O₂ was considered as a weak acid and its hydrogen peroxide anion (OOH^–) can interchange with the hydroxyl group (–OH) of AC generating AC–OOH (Eq. 6). The further reaction between AC–OOH and H₂O₂ will result in the generation of O₂ (Eq. 7) (Georgi and Kopinke, 2005; Rey et al., 2011). The above two pathways of H₂O₂ decomposition may coexist in the AC/H₂O₂ system. However, based on very poor organics degradation and weak ESR signals of active oxygen radicals presented by the AC/H₂O₂ system, it can be inferred that the H₂O₂ decomposition in this system mainly follows a nonradical pathway, which may be useless for organics degradation. Thus, it is suggested that the evaluation of H₂O₂ utilization efficiency should attract more attention when ACs are used as Fenton-like catalysts or catalyst supporters not only in the AC/H₂O₂ system but also in the other H₂O_2⁻involved AOPs (Chen et al., 2023). ${Fe}^{2 +} + H_{2} O_{2} + H^{+} \to {Fe}^{3 +} + • OH + H_{2} O$ (1) ${Fe}^{3 +} + H_{2} O_{2} \to {Fe}^{2 +} + • OOH + H^{+}$ (2) $•OOH \leftrightarrow • O_{2}^{-} + H^{+}$ (3) $AC + H_{2} O_{2} \to {AC}^{+} + {OH}^{-} + • OH$ (4) ${AC}^{+} + H_{2} O_{2} \to AC + H^{+} + • OOH$ (5) $AC - OH + H^{+} {OOH}^{-} \to AC - OOH + H_{2} O$ (6) $AC - OOH + H_{2} O_{2} \to AC - OH + H_{2} O + O_{2}$ (7)

Conclusions

In summary, although the three AC materials prepared with coconut shell, glucose, and corn were able to cause evident H₂O₂ decomposition, the AC/H₂O₂ system presented negligible TC and RhB degradation and weak PNP degradation. The performance of the AC/H₂O₂ system toward organics degradation was much lower than that of the H₂O₂/Fe²⁺ system as the same amount of H₂O₂ consumed, indicating a very low H₂O₂ utilization efficiency in the AC/H₂O₂ system. The poor organics degradation and the weak ESR signals indicated that the H₂O₂ decomposition at AC surface mainly followed a nonradical pathway. According to the results achieved in this study, the role of ACs in the H₂O_2⁻involved AOPs should be considered seriously: Some ACs may be not effective Fenton-like catalysts, and furthermore, the evaluation on H₂O₂ utilization efficiency should attract more attention when ACs are utilized as catalyst supporters.

Footnotes

Authors’ Contributions

H.Z.: Investigation, methodology, and writing—original draft. F.C.: Investigation. Y.C.: Conceptualization, methodology, and writing—review and editing.

Author Disclosure Statement

No competing financial interests exist.

Funding Information

This work was supported by the National Natural Science Foundation of China under Grant 52370079.

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New Insights About the Role of Some Activated Carbon Materials in the H 2 O 2 -Involved Advanced Oxidation of Organic Contaminants

Abstract

Introduction

Experimental Section

Materials

Preparation and characterization of AC materials

Experiments of H2O2 decomposition

Degradation experiments

Electron spin resonance tests

Analytical methods

Results and Discussion

Characterization of various AC materials

H2O2 decomposition caused by various AC materials

TC, RhB, and PNP degradation by AC/H2O2 system and Fenton oxidation system

Mechanism discussion

Conclusions

Footnotes

Authors’ Contributions

Author Disclosure Statement

Funding Information

References

Experiments of H₂O₂ decomposition

H₂O₂ decomposition caused by various AC materials

TC, RhB, and PNP degradation by AC/H₂O₂ system and Fenton oxidation system