Comparative study of atrazine degradation by magnetic clay activated persulfate and H₂O₂

Abstract

To effectively remove the endocrine disrupting chemicals (EDCs) in water, Fe₃O₄ was loaded on the surface of modified sepiolite clay by the method of co-precipitation to catalyze potassium persulfate (K₂S₂O₈) and hydrogen peroxide (H₂O₂) respectively to generate SO₄˙⁻ and ·OH for atrazine (ATZ) removal. The magnetic clay catalyst was characterized by XRD, SEM, N₂ adsorption–desorption and isoelectric point. The degradation efficiency of ATZ in the two systems was systematically compared in terms of initial pH, oxidant dosage and oxidant utilization rate. The results revealed that, after 90 minutes, systems with K₂S₂O₈ and H₂O₂ can remove 65.7% and 57.8% of the ATZ under the given conditions (30 °C, catalyst load: 1 g L⁻¹, initial pH: 5, [ATZ]₀: 10 mg L⁻¹, [H₂O₂]₀: 46 mmol L⁻¹, [PDS]₀: 46 mmol L⁻¹). The magnetic clay catalyst still maintained good catalytic activity and stability during the four consecutive runs. Based on the quenching experiments, it was demonstrated that the dominant radical species in the two systems were SO₄˙⁻/·OH and ·OH, respectively. However, the degradation efficiency of the two systems presented different responses toward the condition variations; the system with K₂S₂O₈ was relatively more sensitive to solution pH, the oxidant efficiency was generally higher than that of the H₂O₂ system (except 184 mmol L⁻¹).

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1. Introduction

Endocrine disrupting chemicals (EDCs), also known as environmental hormones, are contaminants that may cause endocrine disorders. The World Health Organization (WHO) defines it as an exogenous substance or mixture that alters function(s) of the endocrine system and consequently causes adverse health effects in an intact organism, or its progeny, or (sub) populations.¹ Atrazine (ATZ) is a kind of triazine herbicide with toxic properties such as carcinogenicity and endocrine disruption, and has been classed as one of the worst endocrine disruptors. Since 1959, ATZ has been widely used as a chemical herbicide in the world.² Because ATZ has toxic effects on human health, wildlife and the ecological environment, seven European countries including Germany, France and Sweden and the United States have banned the use of ATZ.^3,4 However, due to its structural stability, high solubility, refractory and long-term extensive application, its residues have been detected globally in surface water, groundwater and atmospheric sediments.⁵ Moreover, conventional wastewater technologies, such as biodegradation, coagulation, nanofiltration and sedimentation, cannot remove ATZ effectively in a short time, which makes ATZ degradation more urgent and important.⁶

Advanced oxidation process (AOPs) has been proved to be an effective water treatment technology, which has been widely used in the treatment of refractory organic contaminants in water.^7,8 The traditional AOPs based on ·OH (hydroxyl radical) (HR-AOPs) and the new AOPs based on SO₄˙⁻ (sulfate radical) (SR-AOPs) can produce strong oxidation radicals by catalyzing H₂O₂ and persulfate (PDS), which are used to degrade organic pollutants in water.⁹ Due to the similarity of the structure of H₂O₂ and PDS, there are some general activation modes for them to produce free radicals, such as ultraviolet, ultrasonic, transition metal.^10–13

Fe₃O₄ has an anti-spinel structure, and the electrons can move rapidly between Fe²⁺ and Fe³⁺in octahedron. At the same time, it is a semiconductor with a narrow gap, which is very important for electron transmission. These properties give Fe₃O₄ higher catalytic activity.^14,15 Natural clay is a common catalyst carrier with high specific surface area, good thermal stability, ion-exchange performance and other advantages. When the active metal is supported on the natural clay, the microporous structure of the clay can be improved and the specific surface area can be increased to improve its adsorption and catalytic activity.¹⁶ Natural clay has been widely used in the field of catalysis. Jin et al. loaded Fe₃O₄ onto montmorillonite by co-precipitation, and the phenol (100 mg L⁻¹) was completely removed within 35 min.¹⁷ Huang et al. modified the hydroxy Al–Fe intercalated montmorillonite (FeAl-MT) with CE and La respectively. The results showed that the introduction of CE and LA could form lattice defects and vacancies, which significantly enhanced the catalytic activity of degradation of reactive blue dye.¹⁸ Sepiolite, a kind of magnesium silicate clay mineral with fibrous morphology. It has a standard chemical formula of Mg₈Si₁₂O₃₀(OH)₄(OH₂)₄·8H₂O.¹⁹ Magnetic Fe₃O₄ supported on sepiolite, which can be recovered by external magnetic field due to the magnetism of Fe₃O₄.²⁰

The AOPs based on Fe₃O₄ activation peroxide has been proved to be able to effectively degrade ATZ in water. For example, An et al. used Fe₃O₄ as K₂S₂O₈ activation reagent and UVA/Fe₃O₄/K₂S₂O₈ system to degrade ATZ, the degradation rate of ATZ (100 μmol L⁻¹) reached 90% within 6 h.²¹ Fang used nanometer Fe₃O₄ as catalyst to catalyze H₂O₂, and the degradation rate of ATZ (10 mg L⁻¹) in 48 h was up to 47%.²² In this study, Fe₃O₄ was supported on modified sepiolite by co-precipitation method to prepare magnetic clay catalyst, and its surface properties were characterized. Based on the prepared catalyst, magnetic clay/PDS (PDS system) and magnetic clay/H₂O₂ (H₂O₂ system) were constructed to degrade ATZ in water. The two systems were compared in terms of initial pH, oxidant dosage and oxidant utilization rate.

2. Experimental materials and methods

2.1 Chemicals and instruments

ATZ (99.9%), FeNO₃·9H₂O, FeSO₄·7H₂O, ethanol (EtOH) and tert butyl alcohol (TBA) were purchased from Macklin (Shanghai Macklin Biochemical). Potassium persulfate (K₂S₂O₈), NaOH and HNO₃ were obtained from Sigma Aldrich (St. Louis, USA). Natural sepiolite was purchased from a sepiolite factory in Shijiazhuang. All the chemicals used in this experiment were analytical grade.

Liulian magnetic stirrer (HJ-6, Changzhou Wuyou), pH meter (PHS-3D, Shanghai Leici), UV spectrophotometer (Alpha-1502, Shanghai Puyuan), ultra-performance liquid chromatography (UPLC, 1290 Infinity, Agilent), X-ray diffractometer (XRD, D/max 2550, Rikagu), nitrogen adsorption desorption instrument (SSA-6000, Biaode). Scanning electron microscope (SEM, 1550VP, Zeiss), zeta potentiometer (Zetasizer nano).

2.2 Preparation and characterization of the magnetic clay catalyst

The natural sepiolite was first washed by ultrapure water and then soaked in an acid solution (pH = 2) with solid/water ratio of 1/20 for 2 h to remove the impurities. The solid was then filtered and dried at 80 °C and then heated at 300 °C for another 3 h (magnetic modified sepiolite). The magnetic clay catalyst was prepared by the co-precipitation method. 3.225 g FeNO₃·9H₂O, 1.11 g FeSO₄·7H₂O and 0.8 g modified sepiolite were mixed in the corresponding amount of ultrapure water (200 ml). The solution was stirred by 15 min while being heated in a 60 °C water bath, and then the NaOH (1 mol L⁻¹) was added dropwise until the pH value of the solution reached 9. The obtained dark-brown solution was further heated and stirred for another 30 min, then removed from the water bath pot. After cooling to room temperature, the product was centrifuged and repeatedly rinsed with ultrapure water until the filtrate was neutral. Finally, the mixture was dried overnight at 80 °C to obtain the magnetic clay catalyst.

The crystalline phase of the catalyst was analyzed by XRD (Cu kα radiation, scanning rate 2° min⁻¹, scanning angle 10–90°). The specific surface area and pore size distribution of catalyst were measured by nitrogen adsorption–desorption instrument. The surface morphology of catalysts at different magnification was characterized by SEM. The isoelectric point of catalyst was obtained using zeta potential tester.

2.3 Activity test and quantitative analysis of the ATZ

In each experiment, after adjusting the solution pH with HNO₃ or NaOH, certain amount the magnetic clay catalyst was added in a 100 ml solution of ATZ (10 mg L⁻¹), and the mixture was magnetically stirred at fixed speed for a certain time (30 min) at first to establish adsorption–desorption equilibrium before adding the oxidants (PDS or H₂O₂). At predetermined time intervals, samples were taken by a syringe and immediately filtered by a 0.22 μm polyethersulfone membrane to remove the solid particles, and then be transferred to liquid chromatography injection bottle for testing.

The concentration of ATZ in the solution was detected by UPLC equipped with a UV detector (226 nm for ATZ) using Spheri-5C18 (150 mm × 4.6 mm × 5 μm) column. The mobile phase was composed of 20% ultrapure water and 80% methanol, with a flow rate of 1.0 ml min⁻¹ and an injection volume of 1.0 μL at column temperature of 30 °C.

2.4 PDS concentration analysis

The concentration of PDS was determined by iodimetry.²³ Under neutral condition, PDS reacted with excessive iodine ions (I⁻) to form yellow triiodide, and then the absorbance of the solution was measured at 355 nm by UV-visible spectrophotometer. The PDS content in the solution was calculated according to the relationship between PDS concentration and absorbance.

S₂O₈²⁻ + 3I⁻ → 2SO₄²⁻ + I³⁻ (1)

2.5 H₂O₂ concentration analysis

The concentration of H₂O₂ was determined by titanium oxide sulfate complexation colorimetry.²⁴ In sulfuric acid solution, H₂O₂ reacted with titanium to form an orange complex, and then the absorbance of the solution was measured at 415 nm by UV-visible spectrophotometer. The H₂O₂ content in the solution was calculated according to the relationship between H₂O₂ concentration and absorbance.

TiO²⁺ + H₂O₂ → [TiO(H₂O₂)²⁺] + H₂O₂ (2)

3. Results and discussion

3.1 Characterization of catalyst

3.1.1 Surface morphology analysis. The SEM images of sepiolite support and magnetic clay catalyst are shown in Fig. 1. From Fig. 1 (left), the sepiolite carrier exhibited a fibrous structure with a smooth surface, while the loading of Fe₃O₄ particles (Fig. 1 (right)) made the surface became rough.

Fig. 1 SEM images of sepiolite support (left) and catalyst (right).

3.1.2 Crystal structure analysis. The XRD pattern of the magnetic clay catalyst is presented in Fig. 2. The diffraction peaks at 30°, 35°, 43°, 60° matched well with the face-centered cubic typical phase of Fe₃O₄ (JCPDS card no. 19-0629), which could be attributed to the indices of (220), (311), (511) and (440), respectively, indicating that Fe₃O₄ has been successfully loaded on the sepiolite surface.

Fig. 2 XRD pattern of the magnetic clay catalyst.

3.1.3 Specific surface area and pore size distribution analysis. Fig. 3 depicts the nitrogen adsorption–desorption isotherms and pore size distribution (inset) of the magnetic clay catalyst. It can be seen that the adsorption capacity of catalyst showed an increasing trend with the increase of relative pressure. The adsorption occurred mainly at the relative pressure of 0.8–0.9, implying a characteristic type IV isotherm with a instinct hysteresis loop and catalyst contained mesopores. According to the definition of porous materials, those with pore size less than 2 nm are micropores, those with pore size between 2-50 nm are mesopores, and those with pore size greater than 50 nm are macropores.²⁵ Further, from the inset of Fig. 3, it can be seen that the average pore diameter mainly distributed within the range of 2–50 nm and partly larger than 50 nm, suggesting that the catalyst contained mesopores and macropores. Based on the BET analysis, the specific surface area, average pore size and pore volume of the magnetic clay catalyst were 81.01 m² g⁻¹, 7.88 nm and 0.116 m³ g⁻¹.

Fig. 3 Nitrogen adsorption–desorption isotherms and pore size distribution (inset) of the catalyst.

3.1.4 Isoelectric point analysis. When the solid contacts with the liquid, the phenomenon of charge on the surface of the solid is widespread. The adsorption process of solid particles with oxidants or organics will be affected directly by the charge distribution on the surface of solid particles.²⁶ The point of zero charge (pHzpc) is the pH value of the solution when the particle surface charge is zero.²⁷ This electro-kinetic property can be examined by the change of the potential of the shear layer (i.e. zeta potential) in the diffusion double layer. The isoelectric point of sepiolite and catalyst are shown in Fig. 4, the results illustrate that the pHzpc value of the support and catalyst, which were 7.1 and 7.6 respectively, which indicating that the solid surface is electrically neutral when the solution pH is 7.1 or 7.6 respectively, when it is under 7.1 or 7.6, the surface is positively charged and negatively charged otherwise.

Fig. 4 Isoelectric point of sepiolite and catalyst.

3.2 ATZ degradation in two systems

3.2.1 Activity test. Fig. 5 exhibits the ATZ degradation in different systems. In PDS system (a), after 90 minutes, the catalyst was added alone (1), 11.9% of ATZ was removed at the end of the reaction. When PDS was added alone (2), only 5.5% of ATZ was removed 90 min later, which may be due to the oxidation of persulfate itself. However, in the presence of catalyst/PDS (3), the degradation rate of ATZ increased significantly, reaching to 65.7% after 90 min. Ethanol (EtOH) can react with SO₄˙⁻ and ·OH at the rates constants of 1.6–7.7 × 10⁷ (mol L⁻¹)⁻¹ S⁻¹ and 1.2–2.8 × 10⁹ (mol L⁻¹)⁻¹ S⁻¹, respectively, while tert-butanol (TBA) can react with ·OH at a higher rate of 3.8–7.6 × 10⁸ (mol L⁻¹)⁻¹ S⁻¹, which is approximately 1000 times higher than that of SO₄˙⁻ (4.0–9.1 × 10⁵ (mol L⁻¹)⁻¹ S⁻¹).²⁸ Quenching experiments were conducted adding excess EtOH and TBA to identify and distinguish SO₄˙⁻ and ·OH in two systems. In the system with excessive TBA (4), 60.2% of ATZ was removed, with only 5.5% being cut off, demonstrating that only a small part of the degradation of ATZ was contributed by ·OH, meanwhile, the results of system (5) also confirmed this conclusion. The ATZ degradation was totally inhibited by 52.9% after adding excessive EtOH, confirming that the dominant radicals SO₄˙⁻ of ATZ degradation had been completely quenched by EtOH.

Fig. 5 ATZ 90 min removal efficiency in different systems (%): (a) PDS system (b) H₂O₂ system. (catalyst dose: 1 g L⁻¹, T: 30 °C, pH: 5 oxidant dose: 46 mmol L⁻¹, TBA: 4.6 mol L⁻¹, EtOH: 4.6 mol L⁻¹)²⁹

In the H₂O₂ system (b), 12.3% of ATZ was removed after 90 min when adding catalyst alone (1); only 4.8% of ATZ could be removed after adding only H₂O₂ (2), indicating that H₂O₂ could not generate radicals by itself. After adding catalyst and H₂O₂ (3), it can be seen that the degradation rate of ATZ was significantly promoted to 57.8% after 90 min, which demonstrating that ·OH is the primary radical that was responsible for the ATZ degradation in H₂O₂ system. System (4) also proved this conclusion, the ATZ decomposition was totally terminated by adding EtOH that can annihilate ·OH.

Generally speaking, ATZ could not be effectively degraded by catalyst or oxidant alone because of its refractory property. However, the addition of the magnetic clay catalyst significantly promoted the degradation of ATZ in the two systems, which proved that the catalytic activity of magnetic clay catalyst towards both PDS and H₂O₂ for radicals (SO₄˙⁻ and ·OH), which were responsible for the decomposition of ATZ.

3.2.2 Effects of initial pH on activity. The degradation data of ATZ in the two systems were fitted by quasi first-order kinetics, and the pseudo-first-order rate constant (k_obs) was obtained under different pH value of the solution. Meanwhile, the rate of oxidant after the termination of the reaction (90 min) was measured.

For PDS system (Fig. 6(a)), the corresponding k_obs were 0.0044, 0.0045, 0.0027, 0.0032 and 0.0042 min⁻¹ at solution pH of 3, 5, 7, 9 and 11 respectively. It was obvious that acidic pH is favorable for ATZ degradation. It was owing to two main reasons: (i) organic compounds exist in a neutral state at a pH lower than its pK_a value. Beyond this pK_a value, organic compounds will generate a negative charge.³⁰ The pK_a of ATZ is 1.68, indicating that ATZ was negatively charged in this experiment.³¹ Based on the pHzpc value of the magnetic clay catalyst (Fig. 4), which is 7.6, the catalyst was positively charged at pH < 7.6. Therefore, the electrostatic adsorption between the catalyst surface and ATZ might be enhanced at pH of 3, 5 and 7, and the effective collision efficiency between the catalyst surface active center and S₂O₈²⁻ could also be increased accordingly, so as to generate more radicals that can directly degrade the ATZ adsorbed on the catalyst surface. (ii) The catalyst was negatively charged at pH > 7.6. Thus, the existed electrostatic repulsive-force between catalyst and S₂O₈²⁻ decrease the generation of SO₄˙⁻. Meanwhile, SO₄˙⁻ would also inevitably react with H₂O or OH⁻ to generate ·OH, which has a relatively lower reactivity toward ATZ in basic conditions.^32,33

Fig. 6 Effects of initial pH on ATZ degradation and oxidant consumption (inset): (a) PDS system (b) H₂O₂ system. (catalyst dose: 1 g L⁻¹, T: 30 °C, oxidant dose: 46 mmol L⁻¹).

However, with pH further increased to 11, PDS might be activated by OH⁻ instead of solid catalyst, generating SO₄˙⁻ and superoxide radicals (O₂˙⁻) (eqn (4) and (5)), correspondingly, and ATZ degradation was again enhanced due to the contribution of O₂˙⁻.^34,35 It can be seen from the illustration that at the end of the reaction, PDS consumption rates corresponding to pH 3, 5, 7, 9 and 11 were 58.3%, 57.5%, 49%, 52.5% and 54.5%, respectively.

HO₂⁻ + S₂O₈²⁻ → SO₄˙⁻ + SO₄²⁻ + H⁺ + O₂˙⁻ (5)

The response of H₂O₂ system to initial pH value was different from that of PDS system (Fig. 6(b)). With the increase of solution pH, the k_obs decreased monotonously from 0.0058 min⁻¹ at pH 3 to 0.003 min⁻¹ at pH 7. However, the degradation rate of ATZ was 0 when the pH was 9 or 11. Firstly, in alkaline environment, H₂O₂ tends to generate H₂O and O₂, resulting in no corresponding ATZ degradation.³⁶ Further, because the increase of solution pH inhibits the homolysis of H₂O₂ and then reduces the generation of ·OH, at the same time, the oxidation potential of already generated ·OH in the alkaline condition is far lower than that in the acid environment and is easier to be quenched.³⁷ It can be seen from the illustration that at the end of the reaction, the corresponding H₂O₂ consumption rates at pH 3, 5, 7, 9 and 11 were 64.5%, 54.2%, 20.1, 0 and 0, respectively.

The effect of initial pH value on the two systems was totally different. The maximum ATZ degradation occurred at pH 5, 3 and the corresponding k_obs were 0.045 and 0.0058 min⁻¹ in PDS and H₂O₂ system, respectively. The buffering capacity of PDS system to pH value of solution was significantly higher than that of H₂O₂ system.

3.2.3 Effects of oxidant dosage on activity. By changing the initial dosage of oxidant in the two systems (as shown in Fig. 7), the degradation data of ATZ was fitted with pseudo-first-order kinetics to obtain the k_obs under different oxidant concentrations, and the consumption rate of oxidant after the termination of the reaction (90 min) was measured.

Fig. 7 Effects of oxidant dose on ATZ degradation and oxidant consumption (inset): (a) PDS system (b) H₂O₂ system. (catalyst dose: 1 g L⁻¹, T: 30 °C, pH: 5).

Fig. 7(a) shows the k_obs corresponding to different PDS dosage. As shown in the figure, increasing the PDS concentration from 23 to 184 mM caused an increase of the k_obs by the end of the reaction (90 min) from 0.0039 min⁻¹ to 0.008 min⁻¹, and the corresponding PDS consumption rates were 55%, 58.2%, 65% and 64%, respectively. Suggesting that PDS is necessary to generate radicals and therefore is beneficial for ATZ degradation.

For H₂O₂ system (Fig. 7(b)), after 90 min, the k_obs with H₂O₂ concentrations of 23, 46, 92 and 184 mM, reached the values of 0.0037, 0.040, 0.051 and 0.053 min⁻¹, and the corresponding H₂O₂ consumption rates were 67.2%, 52.2%, 54.3% and 42.7%, respectively. The k_obs increased with the increased of H₂O₂ concentration, indicating that H₂O₂ is the source of active radicals in the system, so it is beneficial to the degradation of ATZ.

Due to the formation rate of SO₄˙⁻ and ·OH was low at low oxidant concentration, resulting the decrease of ATZ degradation. While increasing the oxidant dose, the corresponding k_obs increased owing to the production of SO₄˙⁻ and ·OH raised. However, excessive oxidants may also reversely inhibit ATZ degradation as it can be used as the scavenger of SO₄˙⁻ and ·OH.^38,39 This phenomenon was not observed in this study, because the amount of oxidant was lower than the inhibition point. The k_obs in PDS system was higher than that in H₂O₂ system at the same oxidant dosage.

3.3 Utilization efficiency of oxidant

To evaluate the PDS and H₂O₂ utilization efficiency in two systems, an efficiency indicator (E) was defined by the following formula:

As can be seen from the Tables 1 and 2 that in both systems, the initial pH of the solution and the oxidant dosage affected the utilization efficiency. The PDS system was relatively more sensitive to solution pH. In the alkaline solution, for PDS system, the utilization efficiency didn't change much at pH of 9, and even increased again at pH of 11, which proved the base-activation of PDS. However, values of E in H₂O₂ system decreased sharply, which can be explained by the non-radical decomposition of H₂O₂ under alkaline condition. At pH 7, the high use rate of oxidant in H₂O₂ system might be due to the lower k_obs, more ATZ remaining at the end of the reaction, and the lower consumption rate of oxidant, which ultimately led to the high oxidant utilization. When the dosage of oxidants in the two systems was changed, the oxidant utilization rate in PDS system was generally higher than that in H₂O₂ system (except 184 mM). This may be that the electronegativity of S₂O₈²⁻ produced by PDS dissolving in water can be stronger when the pH value of solution was 5, and the interaction between S₂O₈²⁻, catalyst surface active center and ATZ molecule might be enhanced, so that the generation rate of radicals and corresponding ATZ degradation rate were higher.

Table 1 Oxidant utilization efficiency at initial different solution pH values

pH	3	5	7	9	11
PDS system	0.617	0.609	0.408	0.490	0.569
H2O2 system	0.651	0.572	1.493	—	—

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Table 2 Oxidant utilization rate under different oxidant dosage

Oxidant dose (mmol L^−1)	23	46	92	184
PDS system	0.526	0.604	0.678	0.806
H2O2 system	0.417	0.594	0.681	0.890

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3.4 Reusability and stability of the catalyst

The reusability and stability of the catalyst in two systems were evaluated in terms of ATZ removal by consecutive four batch experiments under the optimal conditions (catalyst dose: 1 g L⁻¹, T: 30 °C, pH: 5, oxidant dose: 46 mmol L⁻¹, t: 90 min). After each reaction, the solid catalyst was separated by external magnetic field from the solution, washed repeatedly by ultrapure water and dried until the next run. As can be seen from Fig. 8, the ATZ removal decreased from 65.7% to 63.3% (fell off by 2.4%) in PDS system, and 57.8% to 53.7% (fell off by 4.1%) in H₂O₂ system, with reaction time of 90 min. The results indicate that although ATZ removal decreased slightly, the recycled magnetic clay catalyst remained reasonable activity during the recycle process.

Fig. 8 ATZ degradation during the recycle runs. (a) PDS system (b) H₂O₂ system. (catalyst dose: 1 g L⁻¹, T: 30 °C, pH: 5, oxidant dose: 46 mmol L⁻¹, t: 90 min).

4. Conclusions

(1) The magnetic clay catalyst was successfully prepared by a co-precipitation method and verified to be an effective heterogeneous catalyst, which can catalyze K₂S₂O₈ and H₂O₂ respectively to generate SO₄˙⁻ and ·OH in the degradation of ATZ. It still maintained good catalytic activity and stability during the four consecutive runs.

(2) SO₄˙⁻/·OHand ·OH were the dominant radicals in PDS system and H₂O₂ system according to the quenching experiments.

(3) The buffer capacity of PDS system was relatively more sensitive to solution pH. The reason might be that on the one hand, PDS hydrolysis produced H⁺, which can buffer the change of pH in solution, and the abundant activation mode for PDS on the other hand. However, in PDS system, the degradation effect of ATZ was best in acidic conditions, and the utilization rate of oxidant was higher than that in alkaline environment, it is possible that the oxidation potential energy of SO₄˙⁻ that played a major role in the acidic environment, is higher than that of O₂˙⁻, which was the primary radical in the alkaline condition.

(4) Both systems were affected by the oxidant dosage, and k_obs increased with increasing oxidant dosage. At the same oxidant concentration, the k_obs value of the PDS system was greater than that of the H₂O₂ system, and the usage rate of the oxidant in the PDS system was generally higher than that of the H₂O₂ system (except 184 mM), indicating that the degradation effect of PDS system on ATZ was obviously better than that of H₂O₂ system. It might be due to the SO₄˙⁻ produced by PDS dissolved in water became electronegative. The effective collision efficiency between the catalyst surface active center and S₂O₈²⁻ was correspondingly increased at pH 5, resulting in more multiple radicals can directly degrade ATZ adsorbed on the catalyst surface, making the ATZ degradation rate higher.

Conflicts of interest

There are no conflicts to declare.

Acknowledgements

This research was funded by the major subject of science and technology of Sichuan province, grant number [No. 2019YFS0509].

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