The effect of Fenton-like oxidation reaction on decolorization efficiency and mineralization of 200 mg/L RhB dye solution at different temperatures was studied. From the acidic (c) and (d) groups with pH 3 and alkaline (e) and (f) groups with pH 10 in Figure 4, the catalyst dosage was 10 g/L and H2O2 concentration remained unchanged at 196 mmol/L. As can be seen from Figure 5, regardless of the system, with the increase in temperature, the oxidation degradation rate was accelerated, RhB removal rate and the removal rate of TOC increased. When pH was 3 and reaction temperature was 35 ◦C, the reaction rate of RhB had an obvious rise after fall, and this trend remained constant. This was because the particle catalyst only had an adsorption effect on RhB at the beginning, or the adsorption rate was much higher than the oxidative degradation rate. With the progress of the reaction, a large amount of H2O2 entered into the iron oxide inside the particle catalyst and on the surface of the particle catalyst to form a heterogeneous Fenton system and generate active free radicals. The RhB molecules adsorbed on the pore surface were oxidized and degraded into small molecules that fall off from the surface, and then continued to adsorb RhB molecules in the solution, forming a cycle of adsorption–degradation–desorption–readsorption–re-degradation–re-desorption. At that time, the oxidation degradation process was dominant, so the reaction rate in the system increased twice. The energy was required for the formation of active free radicals, so the higher the temperature in the system, the higher the energy [54], the greater the probability of breaking the O-O H2O2 bond and the higher the concentration of active free radicals, which significantly accelerated the process of oxidative degradation [55]. Therefore, there was no secondary increase in the reaction rate when the reaction temperature was 55 ◦C. In actual industrial applications such as printing and dyeing, the wastewater discharge temperature is higher, so the reaction temperature of 55 ◦C was chosen for the follow-up experiment.

### *2.3. Study on Sustainability and Stability of the Granular Catalyst*

It is an ideal research target for heterogeneous Fenton catalysts with certain regeneration, reuse and good stability. The particle catalyst reached a saturation state after 360 min. Therefore, H2O2 was added after the particle catalyst was adsorbed for 360 min to investigate the change in the concentration of the RhB dye under the optimal catalytic reaction conditions. As can be seen from Figure 6, only when a granular catalyst was presented in the solution, the RhB residual rate was 50.26% after 360 min. After that, with the increase in adsorption time, the concentration of the RhB dye did not decrease much. At 480 min, the residual rate of RhB was still as high as 47.07%, which indicated that the adsorption of RhB on the granular catalyst reached saturation. Therefore, a certain amount of H2O2 was added immediately after the particle catalyst was adsorbed and saturated. It can be seen that under both acidic and alkaline conditions, the concentration of RhB in the solution

periment.

catalyst.

decreased rapidly, and the residual rate of RhB was less than 1% after 120 min, indicating that after adding H2O2, the particle catalyst could effectively degrade and completely decolorize the RhB molecules adsorbed on the surface of particle channels, so the prepared particle catalyst had a certain in situ regeneration performance. can be seen that under both acidic and alkaline conditions, the concentration of RhB in the solution decreased rapidly, and the residual rate of RhB was less than 1% after 120 min, indicating that after adding H2O2, the particle catalyst could effectively degrade and completely decolorize the RhB molecules adsorbed on the surface of particle channels, so the prepared particle catalyst had a certain in situ regeneration performance.

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*2.3. Study on Sustainability and Stability of the Granular Catalyst*

significantly accelerated the process of oxidative degradation [55]. Therefore, there was no secondary increase in the reaction rate when the reaction temperature was 55 °C. In actual industrial applications such as printing and dyeing, the wastewater discharge temperature is higher, so the reaction temperature of 55 °C was chosen for the follow-up ex-

It is an ideal research target for heterogeneous Fenton catalysts with certain regeneration, reuse and good stability. The particle catalyst reached a saturation state after 360 min. Therefore, H2O2 was added after the particle catalyst was adsorbed for 360 min to investigate the change in the concentration of the RhB dye under the optimal catalytic reaction conditions. As can be seen from Figure 6, only when a granular catalyst was presented in the solution, the RhB residual rate was 50.26% after 360 min. After that, with the increase in adsorption time, the concentration of the RhB dye did not decrease much. At 480 min, the residual rate of RhB was still as high as 47.07%, which indicated that the adsorption of RhB on the granular catalyst reached saturation. Therefore, a certain amount of H2O2 was added immediately after the particle catalyst was adsorbed and saturated. It

**Figure 6.** Concentration change of H2O2 on Rhodamine B after adsorption equilibrium of granular **Figure 6.** Concentration change of H2O2 on Rhodamine B after adsorption equilibrium of granular catalyst.

In practical engineering applications, the catalyst itself must be reusable and have an excellent stable structure [56,57], so the reusability of the annular catalyst is also an important evaluation index. The granular catalyst was put into a 100 mL RhB solution with an initial concentration of 200 mg/L; the catalyst dosage was kept at 10 g/L, the H2O2 concentration was 196 mmol/L, the pH value was 3 and 10, and the solution temperature was 55 °C. The reaction was repeated five times, the granular catalyst was removed and washed then dried with deionized water. As can be seen from Figure 7a,b, RhB can be effectively decolorized and degraded by a granular catalyst 180 min after five repeated In practical engineering applications, the catalyst itself must be reusable and have an excellent stable structure [56,57], so the reusability of the annular catalyst is also an important evaluation index. The granular catalyst was put into a 100 mL RhB solution with an initial concentration of 200 mg/L; the catalyst dosage was kept at 10 g/L, the H2O2 concentration was 196 mmol/L, the pH value was 3 and 10, and the solution temperature was 55 ◦C. The reaction was repeated five times, the granular catalyst was removed and washed then dried with deionized water. As can be seen from Figure 7a,b, RhB can be effectively decolorized and degraded by a granular catalyst 180 min after five repeated reactions under acidic and alkaline conditions. Decoloring performance remained the same, and the basic removal rate was stable and constituted over 98%, which showed that the preparation of catalysts had good reusability. However, the mineralization effect of RhB decreased. After the first use of granular catalyst, the mineralization rate of RhB reached 28.11% (pH = 3) and 25.75% (pH = 10). After repeated reuse, the mineralization rate decreased. The mineralization rate of the fifth use was only 22.16% (pH = 3) and 16.28% (pH = 10), as shown in Figure 7c. This was due to the dissolution of a small amount of infirm bound iron ions on the catalyst's surface after each repeated reuse, as shown in Figure 7d. Another reason was that a small amount of organic matter in the previous reaction was adsorbed and remained on the surface of the catalyst channel, resulting in the reduction of active sites and the TOC removal rate.

### *2.4. Study on the Degradation Effect of Different Dyes by the Granular Catalyst*

The above results showed that the granular catalyst had a significant degradation effect on RhB. The feasibility of its application in dye wastewater can be better clarified through the degradation experiment of different dyes. Methylene blue and Congo red were used as the target pollutants to degrade as RhB in the same reaction system. Figure 8 showed that under alkaline conditions, the reaction rate of cationic RhB was faster than that of the acidic conditions, and the TOC removal rate in the acidic conditions was higher than

that of alkaline conditions. The other cationic dye (methylene blue) also showed a similar trend; the difference was a good effect on the mineralization of the methylene blue on rhodamine B. For cationic dyes, the alkaline condition of the heterogeneous Fenton system was conducive to the oxidation and decolorization of dyes, while the acidic condition was conducive to the degradation and mineralization of dyes. For anionic dyes like Congo red, the decolorization and mineralization effects were worse than those of cationic dyes, which was because the surface of ferric attapulgite granular catalyst prepared in this experiment was partially replaced by Al3+ with Si4+ in the 4-valence coordinates, and by Mg2+ with Al6+ in the 6-times coordinates, resulting in electronegativity on the surface of the granular catalyst. Therefore, it has a certain repulsive effect on anionic dyes. the preparation of catalysts had good reusability. However, the mineralization effect of RhB decreased. After the first use of granular catalyst, the mineralization rate of RhB reached 28.11% (pH = 3) and 25.75% (pH = 10). After repeated reuse, the mineralization rate decreased. The mineralization rate of the fifth use was only 22.16% (pH = 3) and 16.28% (pH = 10), as shown in Figure 7c. This was due to the dissolution of a small amount of infirm bound iron ions on the catalyst's surface after each repeated reuse, as shown in Figure 7d. Another reason was that a small amount of organic matter in the previous reaction was adsorbed and remained on the surface of the catalyst channel, resulting in the reduction of active sites and the TOC removal rate.

same, and the basic removal rate was stable and constituted over 98%, which showed that

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**Figure 7.** Reusability of the granular catalyst under acidic (**a**) and alkaline (**b**) conditions and mineralization rate of Rhodamine B (**c**) and iron dissolution amount (**d**). **Figure 7.** Reusability of the granular catalyst under acidic (**a**) and alkaline (**b**) conditions and mineralization rate of Rhodamine B (**c**) and iron dissolution amount (**d**). *Catalysts* **2022**, *12*, x FOR PEER REVIEW 12 of 21

Congo red, the decolorization and mineralization effects were worse than those of cationic dyes, which was because the surface of ferric attapulgite granular catalyst prepared in this **Figure 8.** Degradation and mineralization of different dyes (0 = 200 mg/L, 22 = 196 mmol/L, T = 55 °C, catalyst = 10 g/L). **Figure 8.** Degradation and mineralization of different dyes (*c*0 = 200 mg/L, *cH*2*O*2 = 196 mmol/L, T = 55 ◦C, catalyst = 10 g/L).

copy. Under the optimal reaction conditions, samples were taken at different times for dilution and placed in a quartz colorimetric dish for full-wavelength scanning within the wavelength range of 250–800 nm, as shown in Figure 9. It can be seen from the figure that the RhB solution had the maximum characteristic absorption peak at 554 nm in the visible region, which was mainly caused by the benzene amino group, carbonyl group, and four ethyl groups in RhB molecular structure [58]. According to relevant literature, with the progress of the reaction process, the degradation of RhB molecule mainly went through n-site diethyl, destruction of the large conjugated structure, ring-opening and mineralization, etc. The macroscopic manifestation was the gradual fading of dye color, and the absorption peak at 554 nm of the UV-VIS spectrum in the figure also gradually weakened. At 180 min, it can be seen that the absorption peak almost disappeared, indicating that the oxidation decolorization of RhB was basically completed and decomposed into small molecule intermediates or completely mineralized into CO2 and H2O. This was due to the similar activity of the Fenton process for the production of oxygen free radicals, which directly attacked the molecular structure of the RhB conjugate with a large variety of anthracene (i.e., benzene amine and carbonyl), which led to its cracking and loss of characteristic absorption peak. At the same time, the hydroxylation effect led to the redshift of the absorption peak, and the N-ethyl process caused a blue shift in its absorption peak [59]. Therefore, it can be concluded that the degradation process of the RhB molecule in this experiment is the simultaneous destruction of n-site diethyl and large conjugated oxanthracene structures, and the red-shift and blue-shift effects cancel each other out, thus

*2.5. Dynamic Study on the Catalytic Degradation Process of the Granular Catalyst*

only showing a decrease in absorbance.

experiment was partially replaced by Al3+ with Si4+ in the 4-valence coordinates, and by Mg2+ with Al6+ in the 6-times coordinates, resulting in electronegativity on the surface of

### *2.5. Dynamic Study on the Catalytic Degradation Process of the Granular Catalyst* 2.5.1. Dynamic UV-VIS Spectra of RhB during the Reaction Process

The degradation process of RhB was described qualitatively by UV-VIS spectroscopy. Under the optimal reaction conditions, samples were taken at different times for dilution and placed in a quartz colorimetric dish for full-wavelength scanning within the wavelength range of 250–800 nm, as shown in Figure 9. It can be seen from the figure that the RhB solution had the maximum characteristic absorption peak at 554 nm in the visible region, which was mainly caused by the benzene amino group, carbonyl group, and four ethyl groups in RhB molecular structure [58]. According to relevant literature, with the progress of the reaction process, the degradation of RhB molecule mainly went through n-site diethyl, destruction of the large conjugated structure, ring-opening and mineralization, etc. The macroscopic manifestation was the gradual fading of dye color, and the absorption peak at 554 nm of the UV-VIS spectrum in the figure also gradually weakened. At 180 min, it can be seen that the absorption peak almost disappeared, indicating that the oxidation decolorization of RhB was basically completed and decomposed into small molecule intermediates or completely mineralized into CO2 and H2O. This was due to the similar activity of the Fenton process for the production of oxygen free radicals, which directly attacked the molecular structure of the RhB conjugate with a large variety of anthracene (i.e., benzene amine and carbonyl), which led to its cracking and loss of characteristic absorption peak. At the same time, the hydroxylation effect led to the redshift of the absorption peak, and the N-ethyl process caused a blue shift in its absorption peak [59]. Therefore, it can be concluded that the degradation process of the RhB molecule in this experiment is the simultaneous destruction of n-site diethyl and large conjugated oxanthracene structures, and the red-shift and blue-shift effects cancel each other out, thus only showing a decrease in absorbance. *Catalysts* **2022**, *12*, x FOR PEER REVIEW 13 of 21

**Figure 9.** UV-vis spectra of RhB degradation (0 = 200 mg/L, cH2O2 = 196 mmol/L, T = 55 °C, catalyst = 10 g/L, pH = 3). **Figure 9.** UV-vis spectra of RhB degradation (*c*0 = 200 mg/L, *cH*2*O*2 = 196 mmol/L, T = 55 ◦C, catalyst = 10 g/L, pH = 3).

2.5.2. Determination of Active Substances in the Reaction Process 2.5.2. Determination of Active Substances in the Reaction Process

In order to investigate whether the prepared granular catalyst was a free radical reaction in catalytic oxidation of RhB, the free radical trapping agent ascorbic acid was added in the reaction process to determine the presence of free radicals in the catalytic system according to the degradation of RhB [60]. Active substances that can catalyze re-In order to investigate whether the prepared granular catalyst was a free radical reaction in catalytic oxidation of RhB, the free radical trapping agent ascorbic acid was added in the reaction process to determine the presence of free radicals in the catalytic system according to the degradation of RhB [60]. Active substances that can catalyze

actions in heterogeneous Fenton systems mainly include hydroxyl radical (•OH) [61], peroxide radical (•OOH) [62], and high-valent iron (FeIV) [63]. To confirm the types of free

tions of tert-butanol (TBA) and potassium iodide were added as hydroxyl radical trapping agents under the optimal reaction conditions [64–66] to determine the existence of hydroxyl radicals in the reaction process. At the same time, a certain concentration of pbenzoquinone was added to the reaction solution to determine whether there were peroxy

As can be seen from Figure 10a, the addition of ascorbic acid led to a gradual decline in the degradation rate of RhB. When the amount of ascorbic acid was 2.5 mmol/L, the residual rate of RhB was almost equal to the adsorption rate of the catalyst, which preliminarily proved that the heterogeneous Fenton reaction system was a free radical reaction. As can be seen from Figure 10b,c, the catalytic degradation rate of the reaction solution with tert-butanol was reduced, indicating that a hydroxyl radical existed in the reaction solution and participated in the catalytic degradation process. However, as the concentration of tert-butanol increased, the degradation of RhB was not completely inhibited. The degradation rate was still 74.97% after 240 min, indicating that there were other active substances in the reaction system. After adding 0.2 mmol/L potassium iodide, the degradation rate of RhB decreased, and the degradation rate of RhB was 79.68% after 180 min. Potassium iodide captured hydroxyl radicals on the surface of the granular catalyst, but its inhibition was relatively small compared with tert-butanol, indicating that the granular catalyst catalyzed the degradation of RhB. The RhB molecules were first adsorbed on the surface of the granular catalyst and then degraded, while hydroxyl radicals catalyzed a small part of RhB molecules in the liquid phase. Figure 10d showed that under the acid condition, the degradation of RhB inhibition was limited, indicating that the heterogeneous Fenton system produced a very little amount of oxygen free radicals, from which it

radicals in the reaction process.

reactions in heterogeneous Fenton systems mainly include hydroxyl radical (•OH) [61], peroxide radical (•OOH) [62], and high-valent iron (FeIV) [63]. To confirm the types of free radicals playing a leading role in the heterogeneous Fenton system, different concentrations of tert-butanol (TBA) and potassium iodide were added as hydroxyl radical trapping agents under the optimal reaction conditions [64–66] to determine the existence of hydroxyl radicals in the reaction process. At the same time, a certain concentration of p-benzoquinone was added to the reaction solution to determine whether there were peroxy radicals in the reaction process.

As can be seen from Figure 10a, the addition of ascorbic acid led to a gradual decline in the degradation rate of RhB. When the amount of ascorbic acid was 2.5 mmol/L, the residual rate of RhB was almost equal to the adsorption rate of the catalyst, which preliminarily proved that the heterogeneous Fenton reaction system was a free radical reaction. As can be seen from Figure 10b,c, the catalytic degradation rate of the reaction solution with tert-butanol was reduced, indicating that a hydroxyl radical existed in the reaction solution and participated in the catalytic degradation process. However, as the concentration of tertbutanol increased, the degradation of RhB was not completely inhibited. The degradation rate was still 74.97% after 240 min, indicating that there were other active substances in the reaction system. After adding 0.2 mmol/L potassium iodide, the degradation rate of RhB decreased, and the degradation rate of RhB was 79.68% after 180 min. Potassium iodide captured hydroxyl radicals on the surface of the granular catalyst, but its inhibition was relatively small compared with tert-butanol, indicating that the granular catalyst catalyzed the degradation of RhB. The RhB molecules were first adsorbed on the surface of the granular catalyst and then degraded, while hydroxyl radicals catalyzed a small part of RhB molecules in the liquid phase. Figure 10d showed that under the acid condition, the degradation of RhB inhibition was limited, indicating that the heterogeneous Fenton system produced a very little amount of oxygen free radicals, from which it can be concluded that the hydroxyl free radical reaction in the process of catalytic degradation of RhB was dominant. *Catalysts* **2022**, *12*, x FOR PEER REVIEW 14 of 21 can be concluded that the hydroxyl free radical reaction in the process of catalytic degradation of RhB was dominant.

**Figure 10.** Effects of different types of free radical trapping agents on degradation of rhodamine B: (**a**) Ascorbic acid; (**b**) TBA; (**c**) KI and TBA; (**d**) P-benzoquinone(c0 = 200 mg/L, 22 = 196 mmol/L, T = 55 °C, catalyst = 10 g/L, pH = 3). **Figure 10.** Effects of different types of free radical trapping agents on degradation of rhodamine B: (**a**) Ascorbic acid; (**b**) TBA; (**c**) KI and TBA; (**d**) P-benzoquinone(*c*0 = 200 mg/L, *cH*2*O*2 = 196 mmol/L, T = 55 ◦C, catalyst = 10 g/L, pH = 3).

2.5.3. Formation Rule of Hydroxyl Radical (•OH) 2.5.3. Formation Rule of Hydroxyl Radical (•OH)

dosage of the catalyst and the hydrogen peroxide concentration.

According to the above analysis, hydroxyl radical reaction plays a dominant role in catalytic degradation of RhB. To investigate the rule of hydroxyl free radical, a certain concentration of coumarin solution was used as the trapping agent of the hydroxyl free radical, and the pH of the solution was 3 under the optimal reaction conditions. The influence on the rule of hydroxyl free radical generation was investigated by changing the According to the above analysis, hydroxyl radical reaction plays a dominant role in catalytic degradation of RhB. To investigate the rule of hydroxyl free radical, a certain concentration of coumarin solution was used as the trapping agent of the hydroxyl free radical, and the pH of the solution was 3 under the optimal reaction conditions. The

rate was accelerated, thus increasing the production of hydroxyl radical.

Figure 11a,b shows that with an increase in the particle catalyst dosage, the genera-

Figure 11c,d showed that with the increase in H2O2 concentration in the solution, the production of hydroxyl radical in the heterogeneous Fenton system also gradually increased, which could be due to the increase in H2O2 concentration under the condition of maintaining the same number of adsorption sites and reactive sites. As a result, more H2O2 could enter the particle catalyst and contact with the iron oxide on the pore surface, increasing the contact frequency (the number of collisions per unit time) and resulting in a faster decomposition rate of H2O2 and production of hydroxyl radicals [67]. However, the generation rule of liquid hydroxyl radicals in the system was slightly different from that of total hydroxyl radicals, as shown in Figure 11e,f. It can be seen that the production amount of liquid hydroxyl radicals was much lower than that of total hydroxyl radicals, indicating that the reaction system was dominated by surface catalytic degradation reaction, supplemented by the liquid radical reaction. With the addition of the catalyst, the amount and rate of formation of liquid hydroxyl radical increased obviously. When the addition of granular catalyst was 1 g/L, the amount of formation of liquid hydroxyl radical

of the particle catalyst which brought more adsorption sites and its reactivity [47] and increased the adsorption quantity of the catalyst. In addition, it increased the contact area of the catalytic reaction. H2O2 was in contact with more iron oxides, and its decomposition

influence on the rule of hydroxyl free radical generation was investigated by changing the dosage of the catalyst and the hydrogen peroxide concentration. *Catalysts* **2022**, *12*, x FOR PEER REVIEW 15 of 21

> Figure 11a,b shows that with an increase in the particle catalyst dosage, the generation of free radicals with similar heterogeneity in the Fenton system gradually increased, and H2O2 consumption also increased. This could be caused by the increase in the dosage of the particle catalyst which brought more adsorption sites and its reactivity [47] and increased the adsorption quantity of the catalyst. In addition, it increased the contact area of the catalytic reaction. H2O2 was in contact with more iron oxides, and its decomposition rate was accelerated, thus increasing the production of hydroxyl radical. reached the maximum. After the addition of the catalyst, the formation rate of liquid hydroxyl radical did not increase. When H2O2 concentration was low, the generation rate of hydroxyl radical increased first and then decreased. With the increase in H2O2 concentration, hydroxyl radical generation in the liquid phase increased first and then decreased, indicating that the higher the H2O2 concentration, the better. A high concentration of H2O2 will lead to the greatest quenching of the hydroxyl radical generated in the liquid phase, reducing the reaction efficiency.

**Figure 11.** Formation rule of hydroxyl radical under different conditions: (**a**,**c**) The consumption of H2O2; (**b**,**d**) •OH production; (**e**) H2O2 concentration (T = 55 °C, catalyst = 10 g/L, pH = 3); (**f**) catalyst dosage (22 = 196 mmol/L, T = 55 °C, pH = 3). **Figure 11.** Formation rule of hydroxyl radical under different conditions: (**a**,**c**) The consumption of H2O2 ; (**b**,**d**) •OH production; (**e**) H2O2 concentration (T = 55 ◦C, catalyst = 10 g/L, pH = 3); (**f**) catalyst dosage (*cH*2*O*2 = 196 mmol/L, T = 55 ◦C, pH = 3).

*2.6. Speculation on the Mechanism of Catalytic Oxidation of RhB Dye by the Granular Catalyst* The degradation mechanism of the heterogeneous Fenton system mainly includes the following three kinds: heterogeneous surface catalysis mechanism, homogeneous catalysis mechanism of iron ion dissolution, and catalytic oxidation mechanism of high iron [68]. The heterogeneous surface catalysis mechanism is as follows: when irradiated with visible light, the dye is excited to reduce part of iron oxide FeIII on the catalyst's surface to FeII, and the catalyst adsorbs the RhB dye and H2O2 on the surface of the granular catalyst. The H2O2 adsorbed on the surface of the granular catalyst reacts with FeII and FeIII on the surface to generate active free radicals. The RhB molecules adsorbed on the catalyst's surface are oxidized and degraded into small intermediate products or partially mineralized into CO2 and H2O. The products are desorbed from the particle surface, and then continue Figure 11c,d showed that with the increase in H2O2 concentration in the solution, the production of hydroxyl radical in the heterogeneous Fenton system also gradually increased, which could be due to the increase in H2O2 concentration under the condition of maintaining the same number of adsorption sites and reactive sites. As a result, more H2O2 could enter the particle catalyst and contact with the iron oxide on the pore surface, increasing the contact frequency (the number of collisions per unit time) and resulting in a faster decomposition rate of H2O2 and production of hydroxyl radicals [67]. However, the generation rule of liquid hydroxyl radicals in the system was slightly different from that of total hydroxyl radicals, as shown in Figure 11e,f. It can be seen that the production amount of liquid hydroxyl radicals was much lower than that of total hydroxyl radicals, indicating that the reaction system was dominated by surface catalytic degradation reaction, supplemented by the liquid radical reaction. With the addition of the catalyst, the amount and rate of formation of liquid hydroxyl radical increased obviously. When the addition of

granular catalyst was 1 g/L, the amount of formation of liquid hydroxyl radical reached the maximum. After the addition of the catalyst, the formation rate of liquid hydroxyl radical did not increase. When H2O2 concentration was low, the generation rate of hydroxyl radical increased first and then decreased. With the increase in H2O2 concentration, hydroxyl radical generation in the liquid phase increased first and then decreased, indicating that the higher the H2O2 concentration, the better. A high concentration of H2O2 will lead to the greatest quenching of the hydroxyl radical generated in the liquid phase, reducing the reaction efficiency.

### *2.6. Speculation on the Mechanism of Catalytic Oxidation of RhB Dye by the Granular Catalyst*

The degradation mechanism of the heterogeneous Fenton system mainly includes the following three kinds: heterogeneous surface catalysis mechanism, homogeneous catalysis mechanism of iron ion dissolution, and catalytic oxidation mechanism of high iron [68]. The heterogeneous surface catalysis mechanism is as follows: when irradiated with visible light, the dye is excited to reduce part of iron oxide FeIII on the catalyst's surface to FeII, and the catalyst adsorbs the RhB dye and H2O2 on the surface of the granular catalyst. The H2O2 adsorbed on the surface of the granular catalyst reacts with FeII and FeIII on the surface to generate active free radicals. The RhB molecules adsorbed on the catalyst's surface are oxidized and degraded into small intermediate products or partially mineralized into CO2 and H2O. The products are desorbed from the particle surface, and then continue to adsorb RhB molecules in the solution, forming a cycle of adsorption–degradation–desorption–readsorption–re-degradation–re-desorption, until RhB is completely degraded. The oxidative degradation process is dominant. The homogeneous catalytic mechanism of iron ion dissolution is as follows: the heterogeneous Fenton particle catalyst of ferric attapulgite dissolves a small number of iron ions during the reaction process, which forms a homogeneous Fenton system with H2O2 in the solution, catalyzes the decomposition of H2O2 to generate active free radicals, and degrades RhB. According to the previous experimental results, the Fenton reaction in this part of dissolved iron ions insignificantly contributes to the degradation of RhB in the system, which is a secondary path of degradation reaction. The mechanism of catalytic oxidation with high iron is as follows: during the conversion of FeII and FeIII on the catalyst surface, some iron oxides are converted into higher FeIV, and high iron directly oxidizes RhB to small molecular intermediates or partially mineralizes into CO2 and H2O [69]. This study took attapulgite particles as the supported and loaded iron ions on the surface of attapulgite to make a heterogeneous granular catalyst, which was used for the degradation of the RhB dye. The mechanism of the degradation process was shown in Figure 12. *Catalysts* **2022**, *12*, x FOR PEER REVIEW 16 of 21 to adsorb RhB molecules in the solution, forming a cycle of adsorption–degradation–desorption–re-adsorption–re-degradation–re-desorption, until RhB is completely degraded. The oxidative degradation process is dominant. The homogeneous catalytic mechanism of iron ion dissolution is as follows: the heterogeneous Fenton particle catalyst of ferric attapulgite dissolves a small number of iron ions during the reaction process, which forms a homogeneous Fenton system with H2O2 in the solution, catalyzes the decomposition of H2O2 to generate active free radicals, and degrades RhB. According to the previous experimental results, the Fenton reaction in this part of dissolved iron ions insignificantly contributes to the degradation of RhB in the system, which is a secondary path of degradation reaction. The mechanism of catalytic oxidation with high iron is as follows: during the conversion of FeII and FeIII on the catalyst surface, some iron oxides are converted into higher FeIV, and high iron directly oxidizes RhB to small molecular intermediates or partially mineralizes into CO2 and H2O [69]. This study took attapulgite particles as the supported and loaded iron ions on the surface of attapulgite to make a heterogeneous granular catalyst, which was used for the degradation of the RhB dye. The mechanism of the degradation process was shown in Figure 12.

**Figure 12.** Proposed catalytic mechanism of dye decolorization and oxidation in the system. **Figure 12.** Proposed catalytic mechanism of dye decolorization and oxidation in the system.

Rhodamine B (RhB; C28H31ClN2O3) was provided by the Tianjin Institute of Chemical

C16H20ClN3S) was provided by Sinopharm Chemical Reagents (Shanghai, China). Sodium hydroxide (NaOH), 30% hydrogen peroxide (H2O2), nitric acid (HNO3), hydrochloric acid (HCI), potassium dichromate (K2Cr2O7), concentrated sulfuric acid (H2SO4), and mercury sulfate (Ag2SO4) were purchased from Nanjing chemical reagents (Nanjing, China). Iron trioxide (Fe2O3), iron nitrate (Fe(NO3)3·9H2O), p-benzoquinone (C6H4O2), potassium iodide (KI), ascorbic acid (C6H8O6), and silver sulfate were provided by Sinopagol Chemical reagents (Nanjing, China). Tert-butyl alcohol (C4H10O) was purchased from Shanghai Lingfeng Chemical reagent (Shanghai, China). Coumarin (C9H6O2) was purchased from Aladdin (Shanghai, China). 7-hydroxycoumarin (C9H6O3) and potassium titanium oxalate (K2TiOC4O8·2H2O) were purchased from McLean (Shanghai, China). All reagents used were analytically pure. Attapulgite (ATP) was found in Xuyi, Jiangsu Province. Atta-

pulgite was prepared into granular supported iron ions as the catalyst.

**3. Experimental Materials and Methods**

*3.1. Reagents and Materials*

### **3. Experimental Materials and Methods**

### *3.1. Reagents and Materials*

Rhodamine B (RhB; C28H31ClN2O3) was provided by the Tianjin Institute of Chemical Reagents (Tianjin, China). Congo red (CR; C32H22N6Na2O6S2) was purchased from Shanghai Maclin Biochemical Technology Co., Ltd. (Shanghai, China). Methylene blue (MB; C16H20ClN3S) was provided by Sinopharm Chemical Reagents (Shanghai, China). Sodium hydroxide (NaOH), 30% hydrogen peroxide (H2O2), nitric acid (HNO3), hydrochloric acid (HCI), potassium dichromate (K2Cr2O7), concentrated sulfuric acid (H2SO4), and mercury sulfate (Ag2SO4) were purchased from Nanjing chemical reagents (Nanjing, China). Iron trioxide (Fe2O3), iron nitrate (Fe(NO3)3·9H2O), p-benzoquinone (C6H4O2), potassium iodide (KI), ascorbic acid (C6H8O6), and silver sulfate were provided by Sinopagol Chemical reagents (Nanjing, China). Tert-butyl alcohol (C4H10O) was purchased from Shanghai Lingfeng Chemical reagent (Shanghai, China). Coumarin (C9H6O2) was purchased from Aladdin (Shanghai, China). 7-hydroxycoumarin (C9H6O3) and potassium titanium oxalate (K2TiOC4O8·2H2O) were purchased from McLean (Shanghai, China). All reagents used were analytically pure. Attapulgite (ATP) was found in Xuyi, Jiangsu Province. Attapulgite was prepared into granular supported iron ions as the catalyst.

### *3.2. Preparation and Characterization of Catalyst*

Different additives and attapulgite were mixed evenly in a certain proportion, and a certain quality of deionized water was added into a molded shape for mixing and rubbed into 1–3 mm particles. The mixture was placed in the oven at 110 ◦C to dry, and then into the tube furnace to calcinate for a certain time. The prepared particles were immersed in a solution of iron ion concentration of 1.6 mol/L, the ratio of solid to liquid was 20:1, the immersion temperature was 80 ◦C, and the immersion time was 4 h. After filtration and repeated washing several times, the obtained products were dried in a 110 ◦C oven and then calcined in a tubular furnace at a certain temperature to prepare an iron-bearing attapulgite granular catalyst.

The ASP-2020 specific surface area and pore size analyzer of Micromeritics were used to measure the specific surface area at the degassing station of the analyzer at a rate of 10 ◦C/min to 300 ◦C for 10 h at the temperature of liquid nitrogen (77 K). EDS analysis was carried out under the conditions of 40–50,000 times magnification and gold spraying on the surface of dry samples. Ultima IV multifunctional composite X-ray diffractometer θ = 5~80◦ (Cu target Kα, λ = 1.5406 A) was used to test the samples; the JCPDS file number was 87–2096. Jsm-7600f Thermal field emission scanning electron microscope (JEOL) was used to spray gold (Au) on the surface of the absolute dry samples with a magnification of 40~50,000 times.

### *3.3. Instruments and Analytical Methods*

The absorbance of rhodamine B in the solution was determined by a TU-1810 UV-vis spectrophotometer at 554 nm and its concentration was analyzed. Meanwhile, the solution's Congo red and methylene blue can be determined by ultraviolet spectrophotometer and visible spectrophotometer at 497 nm and 664 nm. The pH of the solution was measured using an EOTECH pH 700 pH meter. A TOC-vcpn (Shimadzu, Japan) total organic carbon analyzer was used to determine the TOC value of the solution. The concentration of Fe3+ in the reaction solution was determined by a TAS-990 Super atomic absorption spectrophotometer. The reaction of hydrogen peroxide with potassium titanium oxalate was carried out in an acetic acid-sodium acetate buffer solution to form a stable orange complex. The absorbance of this orange complex was measured at 375 nm and the concentration of hydrogen peroxide (H2O2) was analyzed.

### **4. Conclusions**

This study of the use of natural clay minerals content load iron load attapulgite particles of iron catalyst was prepared by using XRD, EDS and BET methods to study the

basic properties of the catalyst particles itself, and found that particle catalysts for RhB dye not only have a catalytic degradation effect, but also has a certain adsorption performance and adhere to the surface of the catalyst particles in an iron oxide crystal shape. In the process of reaction, the amount of iron dissolved in the granular catalyst is very low. In the catalytic reaction process, conditions are changed to investigate their effects on the dye RhB catalytic degradation performance and find out the best reaction conditions. The study found that under the condition of alkaline dye, the decolorization rate is greater than the acid condition; the lower the initial concentration and reaction, the greater the concentration of H2O2 and catalyst dosing quantity, the higher the reaction temperature, and the greater and faster the dye RhB degradation and mineralization. However, pH value has a great influence on the degradation and mineralization of RhB. In terms of the decolorization effect, alkaline condition > acidic condition > neutral condition, and in terms of the mineralization effect, acidic condition > neutral condition > alkaline condition. Under the optimal catalytic reaction conditions, granular catalyst's reusability, stability and degradation effects on different types of dyes were studied. It was found that the prepared granular catalyst has a charge-negative surface, and the degradation and mineralization effects of cationic dyes are better than that of anionic dyes. It was found that hydroxyl radical reaction dominated the catalytic degradation of RhB dye in a heterogeneous Fenton system, in which surface radical reaction was the main catalyst and liquid radical reaction was the auxiliary one. The study shows a simple and environmentally friendly route that can be scaled up and seems to be a promising approach for industrial wastewater treatment in the future.

**Author Contributions:** P.Z.; methodology and review, Z.D.; disgining-analysis and writing, T.L. and X.R.; data curation, M.A.O.; writing—original draft preparation, W.Y., J.H. and H.J.; visualization and collection. All authors have read and agreed to the published version of the manuscript.

**Funding:** This research was funded by the National Key Research and Development Program "Key Technology of Safety Production and Pollution Monitoring of Wood-Based Panel" (2016YFD0600703).

**Institutional Review Board Statement:** Not applicable, this study did not involve human or animal studies.

**Informed Consent Statement:** Not applicable, this study did not involve human or animal studies.

**Data Availability Statement:** This study did not report any data.

**Conflicts of Interest:** The authors declare that they have no conflict of interest.

### **References**

## *Article* **Evaluation of Fe2+/Peracetic Acid to Degrade Three Typical Refractory Pollutants of Textile Wastewater**

**Jiali Yu 1,†, Shihu Shu 1,†, Qiongfang Wang 2 , Naiyun Gao 3 and Yanping Zhu 1,\***

**Abstract:** In this work, the degradation performance of Fe2+/PAA/H2O2 on three typical pollutants (reactive black 5, ANL, and PVA) in textile wastewater was investigated in comparison with Fe2+/H2O2 . Therein, Fe2+/PAA/H2O2 had a high removal on RB5 (99%) mainly owing to the contribution of peroxyl radicals and/or Fe(IV). Fe2+/H2O2 showed a relatively high removal on PVA (28%) mainly resulting from ·OH. Fe2+/PAA/H2O2 and Fe2+/H2O2 showed comparative removals on ANL. Additionally, Fe2+/PAA/H2O2 was more sensitive to pH than Fe2+/H2O2 . The coexisting anions (20–2000 mg/L) showed inhibition on their removals and followed an order of HCO3 − > SO4 2− > Cl−. Humic acid (5 and 10 mg C/L) posed notable inhibition on their removals following an order of reactive black 5 (RB5) > ANL > PVA. In practical wastewater effluent, PVA removal was dramatically inhibited by 88%. Bioluminescent bacteria test results suggested that the toxicity of Fe2+/PAA/H2O2 treated systems was lower than that of Fe2+/H2O2 . RB5 degradation had three possible pathways with the proposed mechanisms of hydroxylation, dehydrogenation, and demethylation. The results may favor the performance evaluation of Fe2+/PAA/H2O2 in the advanced treatment of textile wastewater.

**Keywords:** peracetic acid; advanced oxidation; reactive dyes; aniline; polyvinyl alcohol

### **1. Introduction**

The textile industry was one of the most water-consuming and key industrial branches, especially in developing countries. In China, the amount of textile wastewater, around 80% of which is from the printing and dyeing process [1], ranked third among all the 41 industries and accounted for 10.1% [2]. Generally, the discharged textile wastewater was treated either by on-site treatment plants in the factory or a combination of factories and urban wastewater treatment plants (WWTPs) to meet the wastewater discharge standard. However, the increasingly stringent discharge standard forces the advanced tertiary treatment urgent for the enhanced removal of refractory pollutants.

Reactive dyes, aniline (ANL), and polyvinyl alcohol (PVA) are three types of typical refractory pollutants in printing and dyeing wastewater and are of increasing environmental concern [3]. Reactive dye is the most important dyeing class for cellulosic fibers [4], up to 10–50% of which would flow into wastewater in the dyeing process [5]. ANL is an important intermediate in syntheses of benzidine azo dyes [2] as well as a product derived from the biotransformation of azo dyes [6]. ANL has been listed as a priority pollutant by the Environmental Protection Agency of the United States due to its carcinogenic and mutagenic effects [7]. In the latest amended Discharge Standard of Water Pollutants for Dyeing and Finishing of Textile Industry (GB 4287-2012) in China [8], the discharge limit

**Citation:** Yu, J.; Shu, S.; Wang, Q.; Gao, N.; Zhu, Y. Evaluation of Fe2+/Peracetic Acid to Degrade Three Typical Refractory Pollutants of Textile Wastewater. *Catalysts* **2022**, *12*, 684. https://doi.org/10.3390/ catal12070684

Academic Editors: Hao Xu and Yanbiao Liu

Received: 28 May 2022 Accepted: 14 June 2022 Published: 22 June 2022

**Publisher's Note:** MDPI stays neutral with regard to jurisdictional claims in published maps and institutional affiliations.

**Copyright:** © 2022 by the authors. Licensee MDPI, Basel, Switzerland. This article is an open access article distributed under the terms and conditions of the Creative Commons Attribution (CC BY) license (https:// creativecommons.org/licenses/by/ 4.0/).

of ANLs was regulated as undetected, which was actually difficult for factories to meet at affordable expenses. PVA, a water-soluble refractory polymer, is widely used in the sizing process of cotton blended fabrics and was lost to the effluent during the desizing process [9]. The discharge of PVA may deteriorate the receiving body of water via causing the lack of dissolved oxygen in the aquatic environment and the release of harmful metals from the sediment [10]. The removal of PVA via the conventional biological process was challenging due to its poor biodegradability [11].

The advanced oxidation process (AOP) is widely adopted as the tertiary treatment for the removal of low-level refractory organic pollutants from the secondary effluent of industrial textile wastewater [12–19]. Recently, Fe2+/peracetic acid (PAA) has emerged as a potential alternative to the conventional Fe2+/hydrogen peroxide (H2O2) Fenton AOP [20,21]. PAA has a high disinfection efficiency and less formation of harmful disinfection byproducts (DBPs) compared to those chlorine-based disinfectants [22]. Thus, it is recommended as an attractive disinfectant for secondary and tertiary wastewater effluents in many countries, (e.g., Canada and parts of Europe) [23–25]. In fact, the PAA solution is an equilibrium mixture of PAA, H2O2, and acetic acid [26], and the PAA-based Fenton system was defined as Fe2+/PAA/H2O2. During the Fe2+/PAA/H2O2 process, PAA played a key role within the initial 5 s and H2O2 became the dominant oxidant afterward due to the much higher reaction rate of PAA with Fe2+ (>650 times) compared to that of H2O2 [27]. Activation of PAA by Fe2+ may primarily generate ·OH, CH3C(O)O·, and Fe(IV) according to reactions (1–5) [27,28]. Moreover, ·OH would also react with PAA and H2O2 to generate secondary radicals, (e.g., CH3C(O)·, CH3C(O)OO·, and HO2·) via reactions (6–9), and the reaction rate of ·OH with PAA was much higher compared with H2O2 [27,29]. That said, the proportion of PAA and H2O2 would affect the distribution of organic radicals and ·OH in the Fe2+/PAA/H2O2 process. In previous research [30–33], PAA-based AOP has exhibited structural selectivity in the removal performance of target pollutants and shown comparable or even superior performance compared to H2O2-based AOP.

$$\rm CH\_3C(O)\rm OOH + Fe^{2+} \rightarrow CH\_3C(O)\rm O\cdot + Fe^{3+} + OH^- \tag{1}$$

$$\text{CH}\_3\text{C(O)COOH} + \text{Fe}^{2+} \rightarrow \text{CH}\_3\text{C(O)O}^- + \text{Fe}^{3+} + \cdot\text{OH} \tag{2}$$

$$\rm{CH\_3C(O)COOH} + \rm{Fe^{2+}} \rightarrow \rm{CH\_3C(O)OH} + \rm{Fe^{IV}O^{2+}} \tag{3}$$

$$\rm H\_2O\_2 + Fe^{2+} \rightarrow \cdot OH + Fe^{3+} + OH^- \tag{4}$$

$$\rm H\_2O\_2 + Fe^{2+} \rightarrow H\_2O + Fe^{IV}O^{2+} \tag{5}$$

$$\text{CH}\_3\text{C(O)COOH} + \cdot \text{OH} \rightarrow \text{CH}\_3\text{C(O)CO}\cdot + \text{H}\_2\text{O} \tag{6}$$

$$\cdot \text{CH}\_3\text{C(O)COOH} + \cdot \text{OH} \rightarrow \text{CH}\_3\text{C(O)}\cdot + \text{H}\_2\text{O} + \text{O}\_2\tag{7}$$

$$\text{CH}\_3\text{C(O)COOH} + \text{OH} \rightarrow \text{CH}\_3\text{C(O)OH} + \text{HO}\_2\text{-}\tag{8}$$

$$\cdot \text{H}\_2\text{O}\_2 + \cdot \text{OH} \rightarrow \text{HO}\_2\cdot + \text{H}\_2\text{O} \tag{9}$$

Among the studies about the active dye degradation by PAA-based AOPs, H2O2 and PAA-related organic radicals played dominant roles in the removal of methyl blue and Brilliant Red X-3B, respectively [21,34]. Co(II)-mediated PAA oxidation in previous work has shown a minor contribution of CH3C(O)OO· to the degradation of ANL [30]. In addition, the degradation of PVA has been evaluated mainly in H2O2-based AOPs with the reaction rate constant of PVA with ·OH in the order of 106−107 M−1 s −1 [35], while scarcely in PAA-based AOP. Considering the different structural characteristics of the three typical pollutants, there is a need to comparatively evaluate their degradation efficacies by Fe2+/PAA/H2O2 and assess the contributions of radicals produced, respectively, by PAA and H2O2.

The objectives of this study are to investigate: (1) the effectiveness of Fe2+/PAA/H2O2 to degrade reactive black 5 (RB5), ANL, and PVA in comparison to Fe2+/H2O2; (2) the impact of operating conditions, (i.e., Fe2+ and PAA dosages, coexisting ions and natural organic matter); (3) contributions of ·OH and PAA-related radicals in their degradation under different pH conditions; (4) the acute toxicity alteration during the process; and (5) degradation intermediates and possible pathways. The main novelty of this work is to investigate the relationship between Fe2+/PAA/H2O2 and pollutant structure, obtaining the applicability of the process in textile water treatment.

### **2. Materials and Methods**

### *2.1. Chemicals and Materials*

RB5, ANL, sodium thiosulfate, potassium iodide, acetic acid, ferrous sulfate (FeSO4·7H2O), phosphoric acid, sodium chloride, and sodium sulfate were of analytical grade and obtained from Shanghai Titan Scientific Co., Ltd. (Shanghai, China). Ammonium molybdate ((NH4)2MoO4), sodium hydroxide (NaOH), sulfuric acid, and *p*-chlorobenzoic acid (*p*CBA) were purchased from Sinopharm Chemical Reagent Co. (Shanghai, China). Tert-butanol (TBA), methanol (MeOH), and 5,5-dimethyl-1-pyrrolinr N-oxide (DMPO) were of chromatographic grade and purchased from Fisher Scientific (Fair Lawn, NJ, USA). H2O2 (30%, w.t.) solution, PVA (Type 1788) with an average molecular weight of 46,000 g/mol (hydrolysis degree of 88%), humic acid (HA), N, N-Diethyl-p-phenylenediamine (DPD), and 5,5-dimethyl-1-pyrrolinr N-oxide (DMPO) were purchased from Sigma-Aldrich (Shanghai, China). All solutions were prepared with ultrapure water from a Millipore Milli-Q water system (Direct-Q3 UV). The secondary effluent taken from a municipal wastewater treatment plant (Songjiang District, Shanghai, China) was used as the real wastewater with dissolved organic carbon (DOC) of around 4.5 mg C/L after filtered through a 0.45 µm membrane. PAA solution, containing PAA:H2O2 at a molar ratio of 1.34:1, was freshly prepared according to the reaction (Equation (10)) and stored at 4 ◦C [36].

$$\text{CH}\_3\text{COOH} + \text{H}\_2\text{O}\_2 \overset{\text{H}\_2\text{SO}\_4}{\overset{\text{H}\_2\text{SO}\_4}{\rightleftharpoons}} \text{CH}\_3\text{C(O)}\text{OOH} + \text{H}\_2\text{O} \tag{10}$$

### *2.2. Experimental Procedures*

All experiments were conducted in a 200 mL glass reactor with constant magnetic stirring at room temperature (20 ± 1 ◦C). The reaction solution contained designated concentrations of target pollutants, the initial pH of which was adjusted to 3.0, 4.0, 5.0, 6.0, and 7.0 by sodium hydroxide (1 M) or sulfuric acid (1 M). Reactions were initiated by adding different dosages of PAA and FeSO4·7H2O simultaneously. Samples (1–2.5 mL) were withdrawn within 10 min at predetermined intervals and immediately quenched by excessive sodium thiosulfate ([Na2S2O3]/[PAA]0 molar ratio >10) for the analysis of target compounds. Meanwhile, PAA decay was also monitored by taking samples periodically without adding any quenching agent. To explore the contribution of direct PAA oxidation and the radicals produced from H2O2 contained in PAA solution, additional trials were also conducted by adding PAA only or Fe2+/H2O2 (H2O2 dosage equal to the concentration of H2O2 in PAA solution). Quenching tests were performed by spiking 100 mM TBA or MeOH to the reaction solution before the addition of PAA and Fe2+. The concentrations of TBA or MeOH were high enough to quench reactive radicals. To quantify the steady-state concentration of ·OH in Fe2+/PAA/H2O2 system under different pH conditions, the ·OH probe (*p*CBA) was spiked to the reaction solution and its time-dependent degradation was also analyzed.

The effect of water matrices on the degradation of the three target pollutants in Fe2+/PAA/H2O2 system was assessed by adding Cl− (0–2000 mg/L), SO4 2− (0–2000 mg/L), HCO3 − (0–2000 mg/L), and HA (0–10 mg C/L), respectively, to the reaction solution. The degradation tests were also conducted in practical wastewater effluent. Samples were also taken for oxidized products or DOC analysis at the beginning and end of each test. All experiments were conducted at least in duplicate, and the error bars represented the standard deviation.

### *2.3. Analytical Methods*

The ANL and *p*CBA concentrations were measured by high-performance liquid chromatography (HPLC, Thermo Scientific UltiMate DioNEX 300, Waltham, MA, USA) coupled with a Symmetry-C18 column (5 µm, 4.6 mm × 250 mm) and a UV detector. The mobile phase for ANL was a 65:35 (*v*/*v*) mixture of methanol and ultrapure water at a flow rate of 1 mL/min. The mobile phase for pCBA was a mixture of methanol and phosphoric acid (70:30, *v*/*v* %) with a flow rate of 1 mL/min. The injection volumes of ANL and *p*CBA samples were 10 and 100 µL, respectively. Both ANL and pCBA were analyzed at the wavelength of 230 nm. The RB5 concentration was measured with an ultraviolet spectrophotometer (UV1800) at a wavelength of 598 nm. The PVA concentration was measured using the modified colorimetric method [37]. Briefly, eight 25 mL volumetric flasks were prepared with each containing either 0, 0.05, 0.1, 0.2, 0.3, 0.4, 0.5, and 0.6 mL of 0.5 g/L standard PVA solution and diluted to 10 mL. Then, 5 mL of 4% boric acid and 2 mL of I2-KI (1.27 g/L I2 and 25 g/L of KI) were added. After equilibration for 5 min, the solutions were diluted to 25 mL and measured at a wavelength of 690 nm.

Electron paramagnetic resonance (EPR, EMXnano231, Bruker, Rheinstetten, Germany) was used to determine the reactive species with DMPO as the spin trapping agent, with further details presented in SI. The oxidized products of RB5 by Fe2+/PAA/H2O2 system were analyzed by HPLC-MS (Q Exactive Focus, Thermo Fisher Scientific, Waltham, MA, USA), and ANL was detected by GC-MS (Thermo Fisher Scientific, Waltham, MA, USA), the details of which were provided in SI. The PAA stock solution was regularly calibrated using titration method [7]. Concentrations of PAA and H2O2 in PAA solution were determined according to the Hach DPD method [38]. H2O2 concentration in the absence of PAA was measured using a triiodide absorbance method [39]. Acute ecotoxicity was assessed by the change of bioluminescence intensity bioluminescent with Vibrio fischeri bacteria in toxicity analyzer (HACH, Ames, IA, USA) [40], with the details of the method given in Supplementary Materials.

### **3. Results and Discussions**

### *3.1. Process Degradation Efficiency Assessment*

The degradation behaviors of Fe2+/PAA/H2O2 towards the three pollutants were comparatively evaluated at initial pHs of 3.0, 4.0, 5.0, 6.0, and 7.0 compared with PAA only and Fe2+/H2O2. As shown in Figure 1a, 90% of RB5 was removed through an initial fast degradation (94%) within 5 s, followed by a slow degradation in the Fe2+/PAA/H2O2 system, while a minor RB5 degradation (<5%) occurred in PAA only system due to the slow reaction rate. In the Fe2+/H2O2 system, RB5 removal decreased by 6% in the first 5 s and 75% in the whole process compared with Fe2+/PAA/H2O2, attributed to reactive oxidative species (ROS) generated from PAA in addition to H2O2. By contrast, ANL removal in the Fe2+/PAA/H2O2 system (47%) was comparable to that (39%) in the Fe2+/H2O2 system, which suggested that the degradation efficiency of ANL in the former was mainly attributed to the presence of H2O2 other than PAA. Interestingly, the PVA degradation efficiency was lower in the Fe2+/PAA/H2O2 system compared with the Fe2+/H2O2 system. This was likely because PAA reacted much faster with Fe2+ than H2O2 and thus contributed to its preferential consumption of Fe2+. On the other hand, the presence of PAA may convert ·OH to C− via the reaction of ·OH with PAA according to Equation (10). In comparison with previously reported results (Table 1), this process showed a relatively rapid removal of these three pollutants, especially RB5 [41–46].

Fe2+/PAA/H2O2

**Figure 1.** Time-dependent degradation of (**a**,**d**,**g**,**j**,**m**) RB5, (**b**,**e**,**h**,**k**,**n**) PVA and (**c**,**f**,**i**,**l**,**o**) ANL by different processes in a pH range of 3.0–7.0. Conditions: [PAA]0 = 15 mg/L, [H2O2]0 = 5 mg/L, [RB5]0 = 20 mg/L, [PVA]0 or [ANL]0 = 10 mg/L, initial pH = 3.0, 5.0, and 7.0, T = 21 ± 1 °C. **Figure 1.** Time-dependent degradation of (**a**,**d**,**g**,**j**,**m**) RB5, (**b**,**e**,**h**,**k**,**n**) PVA and (**c**,**f**,**i**,**l**,**o**) ANL by different processes in a pH range of 3.0–7.0. Conditions: [PAA]0 = 15 mg/L, [H2O2 ]0 = 5 mg/L, [RB5]0 = 20 mg/L, [PVA]0 or [ANL]0 = 10 mg/L, initial pH = 3.0, 5.0, and 7.0, T = 21 ± 1 ◦C.

**Table 1.** Comparison of different AOPs in removing RB5, ANL, and PVA with Fe2+/PAA/H2O2 system. **Table 1.** Comparison of different AOPs in removing RB5, ANL, and PVA with Fe2+/PAA/H2O2 system.

RB5 20 10 3.0 1.1 15/5 mg/L 94 This work ANL 10 10 3.0 2.2 47

Fe3O4/PMS RB5 50 60 7.0 250 614.76 mg/L 94.86 [43] AmGO/UV-A RB5 100 120 8.0 5000 40 W 75 [44]

**Table 1.** *Cont*.

Note: PS, persulfate; PMS, peroxymonosulfate; AmGO, amino-Fe3O4-functionalized graphene oxide; SPC, sodium percarbonate; LPH, Litres per hour.

The ·OH was identified in Fe2+/PAA/H2O2 system via EPR. Figure S1 showed that the characteristic peak of the DMPO-HO· spin adduct signal appeared in the spectrum, suggesting the existence of ·OH in the system. To differentiate between the contributions of ·OH and other ROSs, (i.e., peroxyl radicals and Fe(IV)), TBA was used to quench ·OH (*k*·OH/TBA = (3.8–7.6) ×108 M−1 s −1 ) [47], and MeOH was used as a quencher for both ·OH (*k*·OH/MeOH = 9.16×109 M−1 s −1 ) [48] and acetyl(per)oxyl radicals, (i.e., CH3COO· and CH3C(O)OO·) [30]. As shown in Figure 1a–c, TBA significantly inhibited the removal of ANL and PVA by 15% and 17%, respectively, but showed negligible influence on RB5 removal. By contrast, MeOH inhibited the removal of RB5, ANL, and PVA by 25%, 37%, and 17%, respectively. These results further indicated that PVA degradation mainly depended on ·OH, ANL mainly on both ·OH and other ROSs, while RB5 was mainly on other ROSs compared to ·OH.

The influence of pH on the degradation of these three pollutants was investigated in a range of 3.0–7.0. Figure 1 showed that their degradation efficiencies declined with the increase in pH, the extent of which followed an order of RB5 > ANL > PVA. The pH effects for RB5 and PVA were more significant in Fe2+/PAA/H2O2 compared with Fe2+/H2O2, while similar for ANL. These results indicated that other ROSs were more susceptible to pH than ·OH, likely due to the higher reaction rate constants in Equations (1)–(3) (16,000– 110,000 M−1 S −1 ) compared with Equations (4) and (5) (63–76 M−1 S −1 ) [27]. Therein, the former could be inhibited more significantly at the elevated pH mainly because of the higher OH− concentration. Meanwhile, the higher pH could result in a decline in Fe2+ according to reaction 11.

$$\text{Fe}^{2+} + \text{OH}^- \rightarrow \text{Fe(OH)}\_2 \tag{11}$$

### *3.2. Effects of PAA and Fe2+ Dosages on Pollutants' Removal*

The effects of PAA and Fe dosages on the degradation efficiencies of the three pollutants were evaluated. As shown in Figure 2a–c, their removals increased with the PAA dosage rising from 5 to 15 mg/L likely attributed to the increased number of radicals. As the PAA dosage further rose to 30 mg/L, their removals either remained stable (RB5) or decreased, which may be explained by the quenching effect of PAA and/or H2O2 on radicals [30]. These results indicated that the optimal PAA dosage was 15 mg/L.

**Figure 2.** Effects of PAA and Fe2+ dosages on the degradation of (**a**,**d**) RB5, (**b**,**e**) ANL and (**c**,**f**) PVA. Conditions: [RB5]0 = 20 mg/L, [PVA]0 and [ANL]0 = 10 mg/L, initial pH = 3.0, T = 21 ± 1 °C. **Figure 2.** Effects of PAA and Fe2+ dosages on the degradation of (**a**,**d**) RB5, (**b**,**e**) ANL and (**c**,**f**) PVA. Conditions: [RB5]0 = 20 mg/L, [PVA]0 and [ANL]0 = 10 mg/L, initial pH = 3.0, T = 21 ± 1 ◦C.

Moreover, the influence of Fe2+ dosage on their removals was studied with the PAA dosage of 15 mg/L. Figure 2d–f showed that RB5 removal was relatively stable against Fe2+ dosage from 0.55 to 2.20 mg/L, likely owing to the sufficient Fe2+ even at a low dosage (0.55 mg/L) for the activation of PAA/H2O2. ANL removal increased from 11% to 38% with the elevated Fe2+ dosage, implying the enhanced activation of Fe2+ on PAA/H2O2. While PVA removal increased initially (Fe dosage < 2.20 mg/L) and then remained unchanged. These results demonstrated that the dosages of both Fe2+ and PAA/H2O2 had a significant influence on the removal of these pollutants. Moreover, the influence of Fe2+ dosage on their removals was studied with the PAA dosage of 15 mg/L. Figure 2d–f showed that RB5 removal was relatively stable against Fe2+ dosage from 0.55 to 2.20 mg/L, likely owing to the sufficient Fe2+ even at a low dosage (0.55 mg/L) for the activation of PAA/H2O2. ANL removal increased from 11% to 38% with the elevated Fe2+ dosage, implying the enhanced activation of Fe2+ on PAA/H2O2. While PVA removal increased initially (Fe dosage < 2.20 mg/L) and then remained unchanged. These results demonstrated that the dosages of both Fe2+ and PAA/H2O2 had a significant influence on the removal of these pollutants.

### *3.3. Effects of Coexisting Inorganic Anions and Humic Acid on Pollutants' Removal*

*Catalysts* **2022**, *12*, x FOR PEER REVIEW 9 of 15

#### 3.3.1. Effects of Coexisting SO4 2−, Cl−, and HCO3 − *3.3. Effects of Coexisting Inorganic Anions and Humic Acid on Pollutants' Removal*

Considering that textile wastewater generally had high contents of SO4 2−, Cl−, and HCO3 −, their effects on the removal of the three pollutants were investigated with the anions' concentrations of 0–2000 mg/L. Figure 3 demonstrated that all these anions inhibited their removals, the extent of which followed an order of HCO3 − > SO4 2− > Cl−. The inhibition of HCO3 − was probably because of the quenching effect of HCO3 − on ·OH to generate less reactive HCO3 · according to reaction (12) with a high reaction rate constant (>108 M−1 s −1 ) [49]. In addition, HCO3 − may consume Fe2+ to form nonreactive Fe2+-HCO3 − complexes [20], and probably cause a pH increase simultaneously. Among these three pollutants, PVA was the most sensitive against these anions, indicating the possibly easier quenching of ·OH by these anions than other ROSs. As for RB5, SO4 2− at 2000 mg/L or HCO3 − at 200 and 2000 mg/L significantly decreased the removals from 98 to 82.5% or from 98.6 to approximately 81.2%, respectively. The inhibition of SO4 2− and Cl− was likely because they could convert ·OH and peroxyl radicals to SO4· − and chlorine-containing radicals which may show relatively weak oxidative capacity towards these pollutants [36]. 3.3.1. Effects of Coexisting SO42−, Cl−, and HCO3− Considering that textile wastewater generally had high contents of SO42−, Cl−, and HCO3−, their effects on the removal of the three pollutants were investigated with the anions' concentrations of 0–2000 mg/L. Figure 3 demonstrated that all these anions inhibited their removals, the extent of which followed an order of HCO3− > SO42− > Cl−. The inhibition of HCO3− was probably because of the quenching effect of HCO3− on OH to generate less reactive HCO3· according to reaction (12) with a high reaction rate constant (>108 M−1s−1) [49]. In addition, HCO3− may consume Fe2+ to form nonreactive Fe2+-HCO3− complexes [20], and probably cause a pH increase simultaneously. Among these three pollutants, PVA was the most sensitive against these anions, indicating the possibly easier quenching of OH by these anions than other ROSs. As for RB5, SO42− at 2000 mg/L or HCO3− at 200 and 2000 mg/L significantly decreased the removals from 98 to 82.5% or from 98.6 to approximately 81.2%, respectively. The inhibition of SO42− and Cl− was likely because they could convert OH and peroxyl radicals to SO4− and chlorine-containing radicals which may show relatively weak oxidative capacity towards these pollutants [36].

$$\mathrm{HCO\_3^- + \cdot OH \to HCO\_3^- + OH^-} \tag{12}$$

**Figure 3.** Effect of coexisting SO42−, Cl−, and HCO3− on the degradation of (**a**) RB5, (**b**) ANL, and (**c**) PVA in Fe2+/PAA/H2O2 system. Conditions: [PAA] = 15 mg/L, [RB5]0 = 20 mg/L, [PVA]0 or [ANL]0 = 10 mg/L, pH = 3.0, T = 21 ± 1 °C, [anion] = 0, 20, 200, and 2000 mg/L. **Figure 3.** Effect of coexisting SO4 2−, Cl−, and HCO3 − on the degradation of (**a**) RB5, (**b**) ANL, and (**c**) PVA in Fe2+/PAA/H2O2 system. Conditions: [PAA] = 15 mg/L, [RB5]0 = 20 mg/L, [PVA]0 or [ANL]0 = 10 mg/L, pH = 3.0, T = 21 ± 1 ◦C, [anion] = 0, 20, 200, and 2000 mg/L.

#### 3.3.2. Effects of Background HA and Real Water Matrix *Catalysts* **2022**, *12*, x FOR PEER REVIEW 10 of 15

The influence of HA on the degradation of these three pollutants was investigated. Figure 4 showed that HA at a low concentration of 1 mg C/L had a minor effect on their removals, while notable inhibition at 5 and 10 mg C/L. In addition, the inhibition effect followed an order of RB5 > ANL > PVA, indicating a probably higher scavenging capacity of HA on peroxyl radicals compared with ·OH [36]. As for the effect of the water matrix, the removals of RB5, ANL, and PVA decreased from 99%, 35%, and 18% to 56%, 15%, and 7%, respectively. The more pronounced effect on PVA was probably because ·OH, with a non-selective oxidation property, tended to be consumed by background organics in practical wastewater compared to other ROSs. 3.3.2. Effects of Background HA and Real Water Matrix The influence of HA on the degradation of these three pollutants was investigated. Figure 4 showed that HA at a low concentration of 1 mg C/L had a minor effect on their removals, while notable inhibition at 5 and 10 mg C/L. In addition, the inhibition effect followed an order of RB5 > ANL > PVA, indicating a probably higher scavenging capacity of HA on peroxyl radicals compared with OH [36]. As for the effect of the water matrix, the removals of RB5, ANL, and PVA decreased from 99%, 35%, and 18% to 56%, 15%, and 7%, respectively. The more pronounced effect on PVA was probably because OH, with a non-selective oxidation property, tended to be consumed by background organics in practical wastewater compared to other ROSs. 3.3.2. Effects of Background HA and Real Water Matrix The influence of HA on the degradation of these three pollutants was investigated. Figure 4 showed that HA at a low concentration of 1 mg C/L had a minor effect on their removals, while notable inhibition at 5 and 10 mg C/L. In addition, the inhibition effect followed an order of RB5 > ANL > PVA, indicating a probably higher scavenging capacity of HA on peroxyl radicals compared with OH [36]. As for the effect of the water matrix, the removals of RB5, ANL, and PVA decreased from 99%, 35%, and 18% to 56%, 15%, and 7%, respectively. The more pronounced effect on PVA was probably because OH, with a

non-selective oxidation property, tended to be consumed by background organics in prac-

*Catalysts* **2022**, *12*, x FOR PEER REVIEW 10 of 15

**Figure 4.** Effect of coexisting HA and real water matrix on the degradation of RB5, PVA and ANL in Fe2+/PAA/H2O2 system. Conditions: [PAA] = 15 mg/L, [RB5]0 = 20 mg/L, [PVA]0 or [ANL]0 = 10 mg/L, pH = 3.0, T = 21 ± 1 °C. **Figure 4.** Effect of coexisting HA and real water matrix on the degradation of RB5, PVA and ANL in Fe2+/PAA/H2O2 system. Conditions: [PAA] = 15 mg/L, [RB5]0 = 20 mg/L, [PVA]0 or [ANL]0 = 10 mg/L, pH = 3.0, T = 21 ± 1 ◦C. **Figure 4.** Effect of coexisting HA and real water matrix on the degradation of RB5, PVA and ANL in Fe2+/PAA/H2O2 system. Conditions: [PAA] = 15 mg/L, [RB5]0 = 20 mg/L, [PVA]0 or [ANL]0 = 10

#### *3.4. Acute Toxicity Evaluation 3.4. Acute Toxicity Evaluation* mg/L, pH = 3.0, T = 21 ± 1 °C.

A bioluminescent bacteria test was used to evaluate the acute toxicity alteration induced by different AOPs. Figure 5 showed that, after Fe2+/PAA/H2O2 treatment, the toxicity of RB5 and PVA decreased while that of ANL increased. For all these pollutants, the Fe2+/PAA/H2O2 treated systems possessed lower toxicity compared with Fe2+/H2O2 treated ones, implying the eco-friendly advantage of the former. In order to further reduce the toxicity of the effluent, the combination of Fe2+/PAA/H2O2 and subsequent adsorption treatment might be a potential approach. A bioluminescent bacteria test was used to evaluate the acute toxicity alteration induced by different AOPs. Figure 5 showed that, after Fe2+/PAA/H2O2 treatment, the toxicity of RB5 and PVA decreased while that of ANL increased. For all these pollutants, the Fe2+/PAA/H2O2 treated systems possessed lower toxicity compared with Fe2+/H2O2 treated ones, implying the eco-friendly advantage of the former. In order to further reduce the toxicity of the effluent, the combination of Fe2+/PAA/H2O2 and subsequent adsorption treatment might be a potential approach. *3.4. Acute Toxicity Evaluation* A bioluminescent bacteria test was used to evaluate the acute toxicity alteration induced by different AOPs. Figure 5 showed that, after Fe2+/PAA/H2O2 treatment, the toxicity of RB5 and PVA decreased while that of ANL increased. For all these pollutants, the Fe2+/PAA/H2O2 treated systems possessed lower toxicity compared with Fe2+/H2O2 treated ones, implying the eco-friendly advantage of the former. In order to further reduce the toxicity of the effluent, the combination of Fe2+/PAA/H2O2 and subsequent adsorption treatment might be a potential approach.

ditions: [PAA] = 15 mg/L, pH = 3.0, T = 21 ± 1 °C. RB5: ([RB5]0 = 20 mg/L, Fe2+: PAA:H2O2 = 1.1 mg/L:15 mg/L:5 mg/L, Fe2+: H2O2 = 1.1 mg/L:5 mg/L); PVA or ANL: ([PVA]0 or [ANL]0 = 10 mg/L, Fe2+: PAA:H2O2 = 2.2 mg/L:15 mg/L:5 mg/L, Fe2+: H2O2 = 2.2 mg/L:5 mg/L). **Figure 5.** Acute toxicity alteration towards bioluminescent bacteria induced by different AOPs. Conditions: [PAA] = 15 mg/L, pH = 3.0, T = 21 ± 1 °C. RB5: ([RB5]0 = 20 mg/L, Fe2+: PAA:H2O2 = 1.1 mg/L:15 mg/L:5 mg/L, Fe2+: H2O2 = 1.1 mg/L:5 mg/L); PVA or ANL: ([PVA]0 or [ANL]0 = 10 mg/L, Fe2+: PAA:H2O2 = 2.2 mg/L:15 mg/L:5 mg/L, Fe2+: H2O2 = 2.2 mg/L:5 mg/L). **Figure 5.** Acute toxicity alteration towards bioluminescent bacteria induced by different AOPs. Conditions: [PAA] = 15 mg/L, pH = 3.0, T = 21 ± 1 ◦C. RB5: ([RB5]0 = 20 mg/L, Fe2+: PAA:H2O2 = 1.1 mg/L:15 mg/L:5 mg/L, Fe2+: H2O2 = 1.1 mg/L:5 mg/L); PVA or ANL: ([PVA]0 or [ANL]0 = 10 mg/L, Fe2+: PAA:H2O2 = 2.2 mg/L:15 mg/L:5 mg/L, Fe2+: H2O2 = 2.2 mg/L:5 mg/L).

#### *3.5. Intermediate Products and Proposed Pathways 3.5. Intermediate Products and Proposed Pathways 3.5. Intermediate Products and Proposed Pathways 3.5. Intermediate Products and Proposed Pathways 3.5. Intermediate Products and ProposedPathways3.5. Intermediate Products and Proposed Pathways 3.5. Intermediate Products and Proposed Pathways 3.5. Intermediate Products and Proposed Pathways 3.5. Intermediate Products and Proposed Pathways*

Owing to the highest degradation efficiency by Fe2+/PAA/H2O2 among these three pollutants, RB5 was selected as the typical pollutant to identify the degradation products and possible pathways. Nine intermediates were identified (Figure S2) and their structures were proposed in Table 2. Accordingly, Figure 6 exhibited three possible degradation pathways of RB5 (C26H21N5Na4O19S6, *m*/*z* 991) to finally form C10H11N3O3 (*m*/*z* 221), C17H14N3NaO7S2 (*m*/*z* 459), and C10H10N2O (*m*/*z* 174), respectively. Therein, the main mechanisms probably included hydroxylation, dehydrogenation, and demethylation, which was consistent with the previously reported oxidation mechanism related to peroxyl radicals [50,51]. Owing to the highest degradation efficiency by Fe2+/PAA/H2O2 among these three pollutants, RB5 was selected as the typical pollutant to identify the degradation products and possible pathways. Nine intermediates were identified (Figure S2) and their structures were proposed in Table 2. Accordingly, Figure 6 exhibited three possible degradation pathways of RB5 (C26H21N5Na4O19S6, *m/z* 991) to finally form C10H11N3O3 (*m/z* 221), C17H14N3NaO7S2 (*m/z* 459), and C10H10N2O (*m/z* 174), respectively. Therein, the main mechanisms probably included hydroxylation, dehydrogenation, and demethylation, which was consistent with the previously reported oxidation mechanism related to peroxyl radicals [50,51]. Owing to the highest degradation efficiency by Fe2+/PAA/H2O2 among these three pollutants, RB5 was selected as the typical pollutant to identify the degradation products and possible pathways. Nine intermediates were identified (Figure S2) and their structures were proposed in Table 2. Accordingly, Figure 6 exhibited three possible degradation pathways of RB5 (C26H21N5Na4O19S6, *m/z* 991) to finally form C10H11N3O3 (*m/z* 221), C17H14N3NaO7S2 (*m/z* 459), and C10H10N2O (*m/z* 174), respectively. Therein, the main mechanisms probably included hydroxylation, dehydrogenation, and demethylation, which was consistent with the previously reported oxidation mechanism related to peroxyl radicals [50,51]. Owing to the highest degradation efficiency by Fe2+/PAA/H2O2 among these three pollutants, RB5 was selected as the typical pollutant to identify the degradation products and possible pathways. Nine intermediates were identified (Figure S2) and their structures were proposed in Table 2. Accordingly, Figure 6 exhibited three possible degradation pathways of RB5 (C26H21N5Na4O19S6, *m/z* 991) to finally form C10H11N3O3 (*m/z* 221), C17H14N3NaO7S2 (*m/z* 459), and C10H10N2O (*m/z* 174), respectively. Therein, the main mechanisms probably included hydroxylation, dehydrogenation, and demethylation, which was consistent with the previously reported oxidation mechanism related to peroxyl radicals [50,51]. the highest Fe2+/PAA/H2O2RB5 as the pollutant the degradation possible pathways. were (Figure S2) strucin 2. Figure 6 degradafinally form C17H14N3NaO7S2(*m/z* 459), and C10H10N2O (*m/z*174),respectively. Therein, the main mechprobably dehydrogenation,consistent mechanism rad-Owing to the highest degradation efficiency by Fe2+/PAA/H2O2 among these three pollutants, RB5 was selected as the typical pollutant to identify the degradation products and possible pathways. Nine intermediates were identified (Figure S2) and their structures were proposed in Table 2. Accordingly, Figure 6 exhibited three possible degradation pathways of RB5 (C26H21N5Na4O19S6, *m/z* 991) to finally form C10H11N3O3 (*m/z* 221), C17H14N3NaO7S2 (*m/z* 459), and C10H10N2O (*m/z* 174), respectively. Therein, the main mechanisms probably included hydroxylation, dehydrogenation, and demethylation, which was consistent with the previously reported oxidation mechanism related to peroxyl radicals [50,51]. Owing to the highest degradation efficiency by Fe2+/PAA/H2O2 among these three pollutants, RB5 was selected as the typical pollutant to identify the degradation products and possible pathways. Nine intermediates were identified (Figure S2) and their structures were proposed in Table 2. Accordingly, Figure 6 exhibited three possible degradation pathways of RB5 (C26H21N5Na4O19S6, *m/z* 991) to finally form C10H11N3O3 (*m/z* 221), C17H14N3NaO7S2 (*m/z* 459), and C10H10N2O (*m/z* 174), respectively. Therein, the main mechanisms probably included hydroxylation, dehydrogenation, and demethylation, which was consistent with the previously reported oxidation mechanism related to peroxyl radicals [50,51]. Owing to the highest degradation efficiency by Fe2+/PAA/H2O2 among these three pollutants, RB5 was selected as the typical pollutant to identify the degradation products and possible pathways. Nine intermediates were identified (Figure S2) and their structures were proposed in Table 2. Accordingly, Figure 6 exhibited three possible degradation pathways of RB5 (C26H21N5Na4O19S6, *m/z* 991) to finally form C10H11N3O3 (*m/z* 221), C17H14N3NaO7S2 (*m/z* 459), and C10H10N2O (*m/z* 174), respectively. Therein, the main mechanisms probably included hydroxylation, dehydrogenation, and demethylation, which was consistent with the previously reported oxidation mechanism related to peroxyl radicals [50,51]. Owing to the highest degradation efficiency by Fe2+/PAA/H2O2 among these three pollutants, RB5 was selected as the typical pollutant to identify the degradation products and possible pathways. Nine intermediates were identified (Figure S2) and their structures were proposed in Table 2. Accordingly, Figure 6 exhibited three possible degradation pathways of RB5 (C26H21N5Na4O19S6, *m/z* 991) to finally form C10H11N3O3 (*m/z* 221), C17H14N3NaO7S2 (*m/z* 459), and C10H10N2O (*m/z* 174), respectively. Therein, the main mechanisms probably included hydroxylation, dehydrogenation, and demethylation, which was consistent with the previously reported oxidation mechanism related to peroxyl radicals [50,51].

**Table 2.** Details and proposed molecular structure of detected degradation intermediates during Fe2+/PAA/H2O2 oxidation of RB5. **Table 2.** Details and proposed molecular structure of detected degradation intermediates during Fe2+/PAA/H2O2 oxidation of RB5. **Table 2.** Details and proposed molecular structure of detected degradation intermediates during Fe2+/PAA/H2O2 oxidation of RB5. **Table 2.** Details and proposed molecular structure of detected degradation intermediates during Fe2+/PAA/H2O2 oxidation of RB5. icals [50,51].and intermediates RB5.**Table 2.** Details and proposed molecular structure of detected degradation intermediates during Fe2+/PAA/H2O2 oxidation of RB5. **Table 2.** Details and proposed molecular structure of detected degradation intermediates during Fe2+/PAA/H2O2 oxidation of RB5. **Table 2.** Details and proposed molecular structure of detected degradation intermediates during Fe2+/PAA/H2O2 oxidation of RB5. **Table 2.** Details and proposed molecular structure of detected degradation intermediates during Fe2+/PAA/H2O2 oxidation of RB5.

(9)3.71C10H9NO3S223.25224.07

HO3S

HO3S

HO3S

HO3S

HO3S

HO3S

HO3S

HO

NH2

NH2

NH2

NH2

NH2

NH2

NH2

(9) 3.71 C10H9NO3S 223.25 224.07

(9) 3.71 C10H9NO3S 223.25 224.07

(9) 3.71 C10H9NO3S 223.25 224.07

(9) 3.71 C10H9NO3S 223.25 224.07

(9) 3.71 C10H9NO3S 223.25 224.07

(9) 3.71 C10H9NO3S 223.25 224.07

(9) 3.71 C10H9NO3S 223.25 224.07

**No. Retention**

*Catalysts* **2022**, *12*, x FOR PEER REVIEW 11 of 15

Owing to the highest degradation efficiency by Fe2+/PAA/H2O2 among these three pollutants, RB5 was selected as the typical pollutant to identify the degradation products and possible pathways. Nine intermediates were identified (Figure S2) and their structures were proposed in Table 2. Accordingly, Figure 6 exhibited three possible degradation pathways of RB5 (C26H21N5Na4O19S6, *m/z* 991) to finally form C10H11N3O3 (*m/z* 221), C17H14N3NaO7S2 (*m/z* 459), and C10H10N2O (*m/z* 174), respectively. Therein, the main mechanisms probably included hydroxylation, dehydrogenation, and demethylation, which was consistent with the previously reported oxidation mechanism related to peroxyl rad-

**Table 2.**  Intermediate Products and Proposed Pathways*

**Time (min) Chemical Formula Molecular Mass Experimental Mass**

(4) 5.27 C17H14N3NaO7S2 459.43 460.28 OH

**Table 2.** *Cont*.

icals [50,51].

Fe2+/PAA/H2O2 oxidation of RB5.

(1) 5.72 C18H16N3NaO13S4 633.58 634.15

(2) 5.72 C22H15N5Na2O10S3 651.56 652.13

(3) 5.27 C16H11N2NaO7S2 430.39 431.09

(5) 9.94 C26H25N3O8S2 571.62 572.41

(6) 3.71 C26H25N3O7S2 555.62 556.39

**Figure 6.** Proposed degradation pathways of RB5 in Fe2+/PAA/H2O2 system. **Figure 6.** Proposed degradation pathways of RB5 in Fe2+/PAA/H2O2 system. **Figure 6.** Proposed degradation pathways of RB5 in Fe2+/PAA/H2O2 system.

For ANL, six main intermediates were identified with their proposed structures (Table S1) and possible degradation pathways (Figure S3) [52,53]. ANL was firstly attacked by OH to form nitrobenzene (*m/z* 123), N-phenylacetamide (*m/z* 135), and azoxybenzene For ANL, six main intermediates were identified with their proposed structures (Table S1) and possible degradation pathways (Figure S3) [52,53]. ANL was firstly attacked by OH to form nitrobenzene (*m/z* 123), N-phenylacetamide (*m/z* 135), and azoxybenzene (*m/z* 198), all of which could be further degraded to CO2 and H2O. For ANL, six main intermediates were identified with their proposed structures (Table S1) and possible degradation pathways (Figure S3) [52,53]. ANL was firstly attacked by ·OH to form nitrobenzene (*m*/*z* 123), N-phenylacetamide (*m*/*z* 135), and azoxybenzene (*m*/*z* 198), all of which could be further degraded to CO2 and H2O.

#### (*m/z* 198), all of which could be further degraded to CO2 and H2O. **4. Conclusions 4. Conclusions**

www.mdpi.com/xxx/s1.

www.mdpi.com/xxx/s1.

**4. Conclusions** In this work, the degradation of RB5, ANL, and PVA by Fe2+/PAA/H2O2 was investigated compared with Fe2+/H2O2. Therein, Fe2+/PAA/H2O2 and Fe2+/H2O2 were relatively suitable for the degradation of RB5 (94%) and PVA (25%), respectively, while exhibiting similar removal efficiency on ANL. In addition, Fe2+/PAA/H2O2 was more pH-dependent compared with Fe2+/H2O2. Quenching test results indicated that PVA degradation mainly depended on OH, ANL mainly on both OH and other ROSs (peroxyl radicals and Fe(IV)), while RB5 was mainly on other ROSs. Both HCO3− (20–2000 mg/L) and HA (5–10 mg C/L) showed great inhibition in their removals. Among these pollutants, practical effluent showed the greatest inhibition on PVA removal. Toxicity test results demonstrated that, for all these pollutants, Fe2+/PAA/H2O2 treated systems had lower toxicity compared with Fe2+/H2O2 treated ones. Three pathways of RB5 degradation were proposed with the In this work, the degradation of RB5, ANL, and PVA by Fe2+/PAA/H2O2 was investigated compared with Fe2+/H2O2. Therein, Fe2+/PAA/H2O2 and Fe2+/H2O2 were relatively suitable for the degradation of RB5 (94%) and PVA (25%), respectively, while exhibiting similar removal efficiency on ANL. In addition, Fe2+/PAA/H2O2 was more pH-dependent compared with Fe2+/H2O2. Quenching test results indicated that PVA degradation mainly depended on OH, ANL mainly on both OH and other ROSs (peroxyl radicals and Fe(IV)), while RB5 was mainly on other ROSs. Both HCO3− (20–2000 mg/L) and HA (5–10 mg C/L) showed great inhibition in their removals. Among these pollutants, practical effluent showed the greatest inhibition on PVA removal. Toxicity test results demonstrated that, for all these pollutants, Fe2+/PAA/H2O2 treated systems had lower toxicity compared with Fe2+/H2O2 treated ones. Three pathways of RB5 degradation were proposed with the possible mechanisms including hydroxylation, dehydrogenation, and demethylation. In this work, the degradation of RB5, ANL, and PVA by Fe2+/PAA/H2O2 was investigated compared with Fe2+/H2O2. Therein, Fe2+/PAA/H2O2 and Fe2+/H2O2 were relatively suitable for the degradation of RB5 (94%) and PVA (25%), respectively, while exhibiting similar removal efficiency on ANL. In addition, Fe2+/PAA/H2O2 was more pH-dependent compared with Fe2+/H2O2. Quenching test results indicated that PVA degradation mainly depended on ·OH, ANL mainly on both ·OH and other ROSs (peroxyl radicals and Fe(IV)), while RB5 was mainly on other ROSs. Both HCO3 − (20–2000 mg/L) and HA (5–10 mg C/L) showed great inhibition in their removals. Among these pollutants, practical effluent showed the greatest inhibition on PVA removal. Toxicity test results demonstrated that, for all these pollutants, Fe2+/PAA/H2O2 treated systems had lower toxicity compared with Fe2+/H2O2 treated ones. Three pahways of RB5 degradation were proposed with the possible mechanisms including hydroxylation, dehydrogena-

possible mechanisms including hydroxylation, dehydrogenation, and demethylation.

This work may provide guidance to assess the suitability of Fe2+/PAA/H2O2 to efficiently

**Supplementary Materials:** The following supporting information can be downloaded at:

**Author Contributions:** Conceptualization, S.S., Q.W., N.G. and Y.Z.; Data curation, J.Y. and Y.Z.; Funding acquisition, Y.Z.; Investigation, J.Y.; Methodology, Q.W.; Project administration, Y.Z.;

**Supplementary Materials:** The following supporting information can be downloaded at:

**Author Contributions:** Conceptualization, S.S., Q.W., N.G. and Y.Z.; Data curation, J.Y. and Y.Z.; Funding acquisition, Y.Z.; Investigation, J.Y.; Methodology, Q.W.; Project administration, Y.Z.;

remove typical pollutants in textile wastewater.

tion, and demethylation. This work may provide guidance to assess the suitability of Fe2+/PAA/H2O2 to efficiently remove typical pollutants in textile wastewater.

**Supplementary Materials:** The following supporting information can be downloaded at: https: //www.mdpi.com/article/10.3390/catal12070684/s1.

**Author Contributions:** Conceptualization, S.S., Q.W., N.G. and Y.Z.; Data curation, J.Y. and Y.Z.; Funding acquisition, Y.Z.; Investigation, J.Y.; Methodology, Q.W.; Project administration, Y.Z.; Resources, S.S.; Supervision, S.S.; Validation, J.Y.; Writing—original draft, J.Y. and S.S.; Writing review & editing, Q.W., N.G. and Y.Z. All authors have read and agreed to the published version of the manuscript.

**Funding:** This work was supported by the Fundamental Research Funds for Center Universities (21D111311), Shanghai Sailing Program (20YF1401200), the National Natural Science Foundation of China (No. 52000023), and the Shanghai Committee of Science and Technology (No. 19DZ1204400).

**Data Availability Statement:** All data supporting this study are available in the supplementary information accompanying this paper.

**Conflicts of Interest:** The authors declare no conflict of interest.

### **References**

## *Article* **Degradation of Tetracycline Hydrochloride by a Novel CDs/g-C3N4/BiPO4 under Visible-Light Irradiation: Reactivity and Mechanism**

**Wei Qian 1,2,3, Wangtong Hu 2 , Zhifei Jiang 3 , Yongyi Wu 2 , Zihuan Li 2 , Zenghui Diao 2,\* and Mingyu Li 1,\***

**Abstract:** In recent years, with the large-scale use of antibiotics, the pollution of antibiotics in the environment has become increasingly serious and has attracted widespread attention. In this study, a novel CDs/g-C3N4/BiPO4 (CDBPC) composite was successfully synthesized by a hydrothermal method for the removal of the antibiotic tetracycline hydrochloride (TC) in water. The experimental results showed that the synthesized photocatalyst was crystalline rods and cotton balls, accompanied by overlapping layered nanosheet structures, and the specific surface area was as high as 518.50 m2/g. This photocatalyst contains g-C3N4 and bismuth phosphate (BiPO4 ) phases, as well as abundant surface functional groups such as C=N, C-O, and P-O. When the optimal conditions were pH 4, CDBPC dosage of 1 g/L, and TC concentration of 10 mg/L, the degradation rate of TC reached 75.50%. Active species capture experiments showed that the main active species in this photocatalytic system were holes (h+ ), hydroxyl radicals, and superoxide anion radicals. The reaction mechanism for the removal of TC by CDBPC was also proposed. The removal of TC was mainly achieved by the synergy between the adsorption of CDBPC and the oxidation of both holes and hydroxyl radicals. In this system, TC was adsorbed on the surface of CDBPC, and then the adsorbed TC was degraded into small molecular products by an attack with holes and hydroxyl radicals and finally mineralized into carbon dioxide and water. This study indicated that this novel photocatalyst CDBPC has a huge potential for antibiotic removal, which provides a new strategy for antibiotic treatment of wastewater.

**Keywords:** carbon dots; photocatalysis; graphitic carbon nitride; bismuth phosphate; tetracycline hydrochloride

### **1. Introduction**

According to the Chinese Academy of Sciences' antibiotic pollution map, more than 6100 tons of tetracycline (TC) antibiotics are discharged into the soil and groundwater environment every year [1]. Antibiotic concentrations in Chinese rivers generally reach 15 µg/L, far exceeding their toxicity threshold [2,3]. Tetracyclic antibiotics are mainly treated by biological, physical, and chemical oxidation methods [4]. Biological treatment methods mainly rely on microbial decomposition to breakdown pollutants. Tetracycline antibiotics are difficult to completely remove by conventional biological treatment [5,6]. Physical treatment methods include adsorption and membrane treatment technology; however, pollutants are not degraded and can easily cause secondary pollution [7–9]. Chemical oxidation methods mainly degrade various types of pollutants by oxidation and reduction reactions, including ozone oxidation [10], Fenton oxidation [11,12], persulfate oxidation [13,14] and photocatalytic technology [15,16]. Photocatalytic degradation of antibiotics with higher

**Citation:** Qian, W.; Hu, W.; Jiang, Z.; Wu, Y.; Li, Z.; Diao, Z.; Li, M. Degradation of Tetracycline Hydrochloride by a Novel CDs/g-C3N4/BiPO4 under Visible-Light Irradiation: Reactivity and Mechanism. *Catalysts* **2022**, *12*, 774. https://doi.org/10.3390/ catal12070774

Academic Editors: Hao Xu and Yanbiao Liu

Received: 20 June 2022 Accepted: 11 July 2022 Published: 13 July 2022

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efficiency and broader applicability has been studied and applied by a large number of scholars at this stage for the treatment of various antibiotic contaminations [17]. In recent years, graphitic carbon nitride (g-C3N4) has been widely used in photocatalysis. g-C3N4 is a typical conjugated polymer in photocatalysts [18–20]. The conjugated polymer facilitates the transport and separation of photoinduced electrons and holes due to its unique internal conjugated π-bond structure. Additionally, its unique two-dimensional layered structure could assist the transport and migration of electrons, and with the advantages of nontoxicity and visible light response (semiconductor band gap of 2.7 eV), it is widely applied in photocatalytic degradation [21]. However, g-C3N4 suffers from a high photogenerated electron-hole recombination rate, which limits its quantum efficiency [22]. Therefore, promoting the separation of its photogenerated carriers is crucial to promoting its photocatalytic efficiency. It is an effective way to composite different materials to overcome their respective defects. Some researchers have prepared different composites, such as g-C3N4/MoS2 [23], g-C3N4/BiOBr [24], and g-C3N4/BiVO4 [25], which show superior photocatalytic performance. The material composite of BiPO4 and g-C3N4 can compensate for each other's defects and take advantage of the high carrier separation efficiency of BiPO4 and the wide range of light absorption of g-C3N4 [26]. The photocatalytic degradation rate of the g-C3N4/BiPO4 composite photocatalyst was 4.5 times higher than that of P25 TiO2 and 2.5 times higher than that of pure BiPO4 under UV irradiation [27]. This value is 2.5 times higher than that of single mesoporous g-C3N4 under visible light irradiation [28]. Additionally, g-C3N4/BiPO4 composites showed superior catalytic degradation performance in the degradation of TC, SMD, SD, and other pollutants [29–31].

However, the g-C3N4/BiPO4 material could still not absorb sunlight with wavelengths longer than 420 nm. Carbon dots (CDs) have excellent optical properties and can make full use of visible light over 500 nm [32,33]. CDs take advantage of water solubility and biocompatibility. In this paper, we prepared CDs/g-C3N4/BiPO4 composite-based precursors of g-C3N4 and inspected the performance of photocatalytic degradation TC solution, mainly in the following aspects: (a) physicochemical characterization of the CDBPC composite; (b) degradation performance of TC by the CDBPC composite under visible light; (c) degradation kinetics of TC; and (d) proposed reaction mechanism of TC degradation.

### **2. Results and Discussion**

### *2.1. Characterization of CDBPC*

As shown in Figure 1a, the CDBPC catalytic material presented crystal rods and spherical crystals, which agglomerated together and were accompanied by tiny spherical crystals. It was presumed that the crystal rods might be BiPO4. As shown in Figure 1b, the diameter range of crystal rods varied in size, fluctuating in the range of 100–900 nm, while the diameter of spherical crystals was approximately 1 µm. The crystal rods had a smoother surface than the spherical crystals but still had tiny spherical crystals at the edges. In contrast, the surface of the spherical crystals was rougher and more porous. Figure 1c,d presents the transmission electron microscopy scans of the CDBPC composite material. Figure 1c shows that the crystal rods and spherical crystals were tightly wrapped together. Figure 1d shows that the spherical crystal structure was evenly attached to the crystal rods. Therefore, it was presumed that spherical crystals were CD crystals, which was consistent with the results of previous studies [34]. The distribution of elements on the surface of CDBPC material was analyzed by EDS mapping, and the results are shown in Figure 1e–j. The CDBPC sample was mainly composed of five elements, including C, N, O, P, Bi, etc. As shown in the SEM image (Figure 1), the rod-shaped crystals of BiPO4 were mainly concentrated in the southwest corner, while g-C3N4 was mainly distributed in the northeast corner. EDS mapping scans revealed that P, O, and Bi elements were mainly concentrated in the southwest corner and N elements were mainly distributed in the northeast corner, which was consistent with the distribution of BiPO4 and g-C3N4 materials. The C element is evenly distributed on the surface of the material, mainly because the C element consists

materials.

of CDs and g-C3N4. Additionally, CDs exhibited excellent water solubility and were evenly distributed on the surface of materials. element mass ratio accounted for 1.61 wt%, and the Bi mass ratio element accounted for 9.66 wt%.

The semi-quantitative mass ratio and atomic fraction analysis of each element was carried out by EDS, and the results were shown in Table 1. Accordingly, the C element accounted mass ratio of 34.84 wt%, N element mass ratio accounted for 27.98 wt%, the P

C3N4 was mainly distributed in the northeast corner. EDS mapping scans revealed that P, O, and Bi elements were mainly concentrated in the southwest corner and N elements were mainly distributed in the northeast corner, which was consistent with the distribution of BiPO4 and g-C3N4 materials. The C element is evenly distributed on the surface of the material, mainly because the C element consists of CDs and g-C3N4. Additionally, CDs exhibited excellent water solubility and were evenly distributed on the surface of

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**Figure 1.**   Elemental composition of CDBPC composite. **Chemical Elements C N Bi P O**  mass ratio (wt %) 34.84 37.98 15.91 1.61 9.66 atomic fraction (%) 43.26 40.45 0.69 0.78 14.83 The semi-quantitative mass ratio and atomic fraction analysis of each element was carried out by EDS, and the results were shown in Table 1. Accordingly, the C element accounted mass ratio of 34.84 wt%, N element mass ratio accounted for 27.98 wt%, the P element mass ratio accounted for 1.61 wt%, and the Bi mass ratio element accounted for 9.66 wt%.

**Table 1.** Elemental composition of CDBPC composite.

As seen in Figure 2b, the adsorption–desorption curve of the CDBPC material was type IV, indicating the presence of a homogeneous mesoporous structure in the material, which was multilayered. Therefore, it caused the hysteresis loops. The curve belonged to the H3

type. The hysteresis loop was not closed, probably because the adsorption equilibrium was not reached when the relative pressure was close to the saturation vapor pressure [35]. This result indicated that the CDBPC materials had narrow and long pore structures. The jump in the curve of the adsorption isotherm at a relative pressure of 0.1 could illustrate the presence of microporosity in the sample material. From Figure 2a, CDBPC was a mesoporous material with a high pore volume. When the average pore size was between 25 and 5 nm, the pore volume accumulated to 0.01 cm3/g, and the average pore size was ≥5 nm. The cumulative rate of pore volume increased when the average pore size was between 5 and 2.5 nm, reaching a maximum of 0.09 cm3/g cumulatively at an average pore size ≥2.5 nm. The pore-specific surface area reached a maximum value of 162 m2/g when the average pore size was ≥2.5 nm. After calculation by the BET method, the specific surface area of the CDBPC material reached 518.4978 m2/g with an average pore diameter of 2.14 nm and a pore volume of 0.2774 cm3/g. the H3 type. The hysteresis loop was not closed, probably because the adsorption equilibrium was not reached when the relative pressure was close to the saturation vapor pressure [35]. This result indicated that the CDBPC materials had narrow and long pore structures. The jump in the curve of the adsorption isotherm at a relative pressure of 0.1 could illustrate the presence of microporosity in the sample material. From Figure 2a, CDBPC was a mesoporous material with a high pore volume. When the average pore size was between 25 and 5 nm, the pore volume accumulated to 0.01 cm3/g, and the average pore size was ≥5 nm. The cumulative rate of pore volume increased when the average pore size was between 5 and 2.5 nm, reaching a maximum of 0.09 cm3/g cumulatively at an average pore size ≥2.5 nm. The pore-specific surface area reached a maximum value of 162 m2/g when the average pore size was ≥2.5 nm. After calculation by the BET method, the specific surface area of the CDBPC material reached 518.4978 m2/g with an average pore diameter of 2.14 nm and a pore volume of 0.2774 cm3/g.

As seen in Figure 2b, the adsorption–desorption curve of the CDBPC material was type IV, indicating the presence of a homogeneous mesoporous structure in the material, which was multilayered. Therefore, it caused the hysteresis loops. The curve belonged to

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**Figure 2.** Pore size and volume accumulation curves (**a**) and (**b**) nitrogen adsorption–desorption isotherm of CDBPC. **Figure 2.** Pore size and volume accumulation curves (**a**) and (**b**) nitrogen adsorption–desorption isotherm of CDBPC.

As shown in Figure 3, CDBPC exhibited the peaks at 17.02°, 19.18°, 21.50°, 27.32°, 29.00°, 31.33°, 34.62°, 36.84°, 41.92°, 52.80°, 56.68°, 61.22° and 70.91°, corresponding to (101), (011), (111), (200), (120), (200), (−212), (202), (−103), (−402), (132), (233) and (−134) of BiPO4 (JCPDS 80-0209), respectively. The lattice spacing of d(120) is 0.3706 and was calculated by Bragg's diffraction law, which is confirmed by the HRTEM reported in the study [10]. Our calculated value was in very good approximation to standard values. The diffraction peak of CDBPC materials at 26.1° was the characteristic peak of hexagonal CDs (JCPDS 87-1523). CDBPC had broad peaks at 13.0°, 47.62°, and 53.66°, which were characteristic of g-C3N4 (JCPDS 87-1524, 87-1526). The diffraction peaks of BiPO4 and g-C3N4 were sharp and intense, indicating their highly crystalline nature. No impurity peaks were observed, confirming the high purity of the products. From the number and area of peaks detected by XRD, it was shown that BiPO4, followed by g-C3N4 and CDs, was the main component, which was consistent with the results of previous studies As shown in Figure 3, CDBPC exhibited the peaks at 17.02◦ , 19.18◦ , 21.50◦ , 27.32◦ , 29.00◦ , 31.33◦ , 34.62◦ , 36.84◦ , 41.92◦ , 52.80◦ , 56.68◦ , 61.22◦ and 70.91◦ , corresponding to (101), (011), (111), (200), (120), (200), (−212), (202), (−103), (−402), (132), (233) and (−134) of BiPO4 (JCPDS 80-0209), respectively. The lattice spacing of d(120) is 0.3706 and was calculated by Bragg's diffraction law, which is confirmed by the HRTEM reported in the study [10]. Our calculated value was in very good approximation to standard values. The diffraction peak of CDBPC materials at 26.1◦ was the characteristic peak of hexagonal CDs (JCPDS 87-1523). CDBPC had broad peaks at 13.0◦ , 47.62◦ , and 53.66◦ , which were characteristic of g-C3N4 (JCPDS 87-1524, 87-1526). The diffraction peaks of BiPO4 and g-C3N4 were sharp and intense, indicating their highly crystalline nature. No impurity peaks were observed, confirming the high purity of the products. From the number and area of peaks detected by XRD, it was shown that BiPO4, followed by g-C3N4 and CDs, was the main component, which was consistent with the results of previous studies [36,37].

[36,37].

**Figure 3.** XRD patterns of CDBPC. **Figure 3.** XRD patterns of CDBPC.

The CDBPC material was analyzed by Fourier transform infrared spectrograms, and the results are shown in Figure 4. The peaks in the wavenumber range between 680 and 480 cm−1 were caused by bending vibrations of BiPO4, while the peaks near 560 cm−1 were due to O-P-O bonding vibrations. A sharp peak at approximately 1100 cm−1 with a small and not obvious width was due to the stretching vibration of C-O, and the adsorption peak at 1400 cm−1 resulted from C-OH stretching vibrations, which was consistent with previous studies [38]. The characteristic peak near 1500 cm−1 was attributed to the C=N stretching vibration in the carbon–nitrogen heterocycle, which was in agreement with other studies [39]. The spike near 1700 cm−1 with a small width belonged to the C=O stretching vibration of the carboxylic groups, which meant that carboxylic groups were present on the surface of the composite material. There was also a relatively wide peak near 3400 cm−1, and such a peak was mainly caused by C-H bending; another weaker The CDBPC material was analyzed by Fourier transform infrared spectrograms, and the results are shown in Figure 4. The peaks in the wavenumber range between 680 and 480 cm−1 were caused by bending vibrations of BiPO4, while the peaks near 560 cm−1 were due to O-P-O bonding vibrations. A sharp peak at approximately 1100 cm−1 with a small and not obvious width was due to the stretching vibration of C-O, and the adsorption peak at 1400 cm−1 resulted from C-OH stretching vibrations, which was consistent with previous studies [38]. The characteristic peak near 1500 cm−1 was attributed to the C=N stretching vibration in the carbon–nitrogen heterocycle, which was in agreement with other studies [39]. The spike near 1700 cm−1 with a small width belonged to the C=O stretching vibration of the carboxylic groups, which meant that carboxylic groups were present on the surface of the composite material. There was also a relatively wide peak near 3400 cm−1 , and such a peak was mainly caused by C-H bending; another weaker and relatively wide peak near 3500 cm−1 was due to the bending vibration of -OH.

and relatively wide peak near 3500 cm−1 was due to the bending vibration of -OH. As seen in Figure 5a, both CDBPC and BiPO4/g-C3N4 had obvious absorption band edges at 280 nm and 410 nm. Compared with BiPO4/g-C3N4, the absorption intensity of CDBPC decreased, but its absorption range increased significantly. A redshift in the absorption band edge of CDBPC implied that the addition of g-C3N4 reduced the band gap of the photocatalyst. The energy band gap of the two materials was calculated by a Tauc plot, and the results are shown in Figure 5b. The energy band gap was 2.98 eV for BiPO4/g-C3N4, 2.68 eV for the CDBPC material, and 2.98 eV for BiPO4/g-C3N4, indicating that the addition of CDs reduced the energy band gap and resulted in a wider light absorption range.

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**Figure 4.** FT−IR spectra of the CDBPC composite. **Figure 4.** FT−IR spectra of the CDBPC composite. sorption range.

**Figure 5.** (**a**) UV–Vis absorption spectra, (**b**) Tauc plots curve of CDBPC. **Figure 5.** (**a**) UV–Vis absorption spectra, (**b**) Tauc plots curve of CDBPC.

#### *2.2. Effect of Preparation Conditions on CDBPC Catalytic Performance 2.2. Effect of Preparation Conditions on CDBPC Catalytic Performance*

**Figure 5.** (**a**) UV–Vis absorption spectra, (**b**) Tauc plots curve of CDBPC. *2.2. Effect of Preparation Conditions on CDBPC Catalytic Performance*  The effect of CD loading on the catalytic efficiency of TC degradation was investigated, with CD loading ranging from 3% to 30%, and the results are shown in Figure 6. After 60 min of light avoidance adsorption, the highest removal rate of TC was only 4%, which indicated that the adsorption performance of the CDBPC material was poor. Dur-The effect of CD loading on the catalytic efficiency of TC degradation was investigated, with CD loading ranging from 3% to 30%, and the results are shown in Figure 6. After 60 min of light avoidance adsorption, the highest removal rate of TC was only 4%, which indicated that the adsorption performance of the CDBPC material was poor. Dur-The effect of CD loading on the catalytic efficiency of TC degradation was investigated, with CD loading ranging from 3% to 30%, and the results are shown in Figure 6. After 60 min of light avoidance adsorption, the highest removal rate of TC was only 4%, which indicated that the adsorption performance of the CDBPC material was poor. During the photocatalytic reaction, the degradation rate of TC gradually improved with increasing time and became stable at 210 min. The catalytic degradation efficiency was 45.5%. When the CDs loading was 3 wt%, the catalytic degradation efficiency was 48.5% at 210 min, and the catalytic degradation efficiency reached 68.86% when the CDs loading was 5 wt%. Subsequently, when the CDs loading was increased to 10 wt% and 20 wt%, the degradation effect decreased slightly. When the loading of CDs was raised to 30 wt%, the removal efficiency dropped dramatically, and the final degradation rate was only 27.27%. Firstly, due to the excessive CDs covering the g-C3N4 and BiPO4 materials, the two materials could not be excited by light to produce electron holes, and only the light energy that penetrated the CDs could have an effect [39]. Secondly, it was possible that the excessive CDs reduced the number of active sites in the CDs/g-C3N4/BiPO4 composites, which reduced the

effective collision probability between the composites and TC [40]. Thus, it was observed that increasing the loading of CDs appropriately was beneficial to improve the degradation efficiency of TC. However, excessive loading was not conducive to degradation. Under the experimental conditions of this paper, the optimal CDs loading was 5 wt%. cial to improve the degradation efficiency of TC. However, excessive loading was not conducive to degradation. Under the experimental conditions of this paper, the optimal CDs loading was 5 wt%.

ing the photocatalytic reaction, the degradation rate of TC gradually improved with increasing time and became stable at 210 min. The catalytic degradation efficiency was 45.5%. When the CDs loading was 3 wt%, the catalytic degradation efficiency was 48.5% at 210 min, and the catalytic degradation efficiency reached 68.86% when the CDs loading was 5 wt%. Subsequently, when the CDs loading was increased to 10 wt% and 20 wt%, the degradation effect decreased slightly. When the loading of CDs was raised to 30 wt%, the removal efficiency dropped dramatically, and the final degradation rate was only 27.27%. Firstly, due to the excessive CDs covering the g-C3N4 and BiPO4 materials, the two materials could not be excited by light to produce electron holes, and only the light energy that penetrated the CDs could have an effect [39]. Secondly, it was possible that the excessive CDs reduced the number of active sites in the CDs/g-C3N4/BiPO4 composites, which reduced the effective collision probability between the composites and TC [40]. Thus, it was observed that increasing the loading of CDs appropriately was benefi-

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**Figure 6.** Degradation efficiency of TC versus time with different CD quality fractions. **Figure 6.** Degradation efficiency of TC versus time with different CD quality fractions.

#### *2.3. Effects of Operating Parameters on the Degradation of TC 2.3. Effects of Operating Parameters on the Degradation of TC*

The pH had a strong impact on the photocatalytic degradation process. It determined the surface charge properties of the composite, affected the adsorption and degradation ability of the composite on the target pollutant, and determined the ionized state of the catalyst surface. In this paper, the effect of pH on the degradation rate of CDBPC was investigated, and the results are shown in Figure 7a. According to the figure, the pH variation had little effect on the adsorption efficiency during the lightavoidance adsorption stage, which was maintained at a low level. In the photocatalytic degradation stage, the pH was in the range of 4–14, and the degradation rate decreased gradually from 74.89% to 13.67% with increasing pH. When the pH was 2, the degradation rate of TC was 75.02%, which was close to that at pH 4. This was mainly caused by two reasons: (a) acidic conditions could inhibit the decomposition of −OH and the oxygen evolution reaction and improve the utilization of free radicals; (b) compared with The pH had a strong impact on the photocatalytic degradation process. It determined the surface charge properties of the composite, affected the adsorption and degradation ability of the composite on the target pollutant, and determined the ionized state of the catalyst surface. In this paper, the effect of pH on the degradation rate of CDBPC was investigated, and the results are shown in Figure 7a. According to the figure, the pH variation had little effect on the adsorption efficiency during the light-avoidance adsorption stage, which was maintained at a low level. In the photocatalytic degradation stage, the pH was in the range of 4–14, and the degradation rate decreased gradually from 74.89% to 13.67% with increasing pH. When the pH was 2, the degradation rate of TC was 75.02%, which was close to that at pH 4. This was mainly caused by two reasons: (a) acidic conditions could inhibit the decomposition of −OH and the oxygen evolution reaction and improve the utilization of free radicals; (b) compared with the −OH oxidation potential under alkaline conditions (2.02 V), the −OH oxidation potential under acidic conditions (2.85 V) exhibited stronger oxidation performance. After comprehensive consideration, the optimal pH value was 4. In general, the acidic conditions were more favorable for the photocatalytic degradation of TC by CDBPC.

The effect of CBPC dosage on the degradation efficiency of TC was studied, and the results are shown in Figure 7b. When the doses of CDBPC were 0.1 g/L, 0.2 g/L, 0.5 g/L and 1 g/L, the degradation rates of TC were 39.6%, 49.3%, 69.5% and 73.2%, respectively. At this time, as the dosage increased, the degradation rate increased, mainly because the addition of a catalyst could enhance the chance of contact between the catalyst and solute, thereby improving the degradation efficiency. As the dosage continued to increase to 2 g/L and 3 g/L, the degradation rate of TC not only did not improve but also decreased slightly to 71.4% and 71.7%, respectively. This might be due to the high dosage of CDBPC, which caused the materials to block the light from each other and reduced the production of active

species. A large number of catalysts might agglomerate to hinder the contact between the materials and the contaminants.

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**Figure 7.** Degradation efficiency of TC versus time with (**a**) different initial pH values, (**b**) doses of CDBPC composite, and (**c**) initial TC concentrations. **Figure 7.** Degradation efficiency of TC versus time with (**a**) different initial pH values, (**b**) doses of CDBPC composite, and (**c**) initial TC concentrations.

*2.4. Degradation Kinetics of TC*  TC degradation was fitted to pseudo-first-order kinetics and pseudo-second-order kinetics models, respectively (Figure 8). The goodness of (R2), slope (k), and half-lives (t1/2) of the kinetic equations for the different TC initial concentrations and different ini-Figure 7c shows the effect of the initial concentration of TC solution on the photocatalytic degradation rate. In general, the initial concentration of TC had little effect on the light-avoiding adsorption stage. In the photocatalytic stage, when the concentration of TC was 5 mg/L, 10 mg/L, 20 mg/L, 30 mg/L, 50 mg/L, and 100 mg/L, the degradation rates were 79.3%, 75.5%, 69.0%, 51.2%, 34.9%, and 19.7%, respectively. As the initial

tial pH values were calculated in Table 2. For different initial TC concentration conditions, the R2 values of the pseudo-first-order kinetics model ranged from 0.9314 to

0.9949. This suggests that TC degradation was more consistent with the pseudo-secondorder kinetics model. As shown in Figure 8c,d, there was a significant difference in the degradation effect under different initial pH conditions. The fitting linear correlation coefficient was only 0.63–0.74 when the pH value was greater than 7. In the range of pH 2–

**Initial Concentration (mg/L)**

concentration increased, the degradation rate gradually decreased, while the amount of TC degradation increased instead. Mainly due to the high initial concentration of the TC solution, the catalyst had more contact opportunities with TC, and thus, the amount of TC degradation was greater. Similarly, because of the high initial concentration of TC solution, the degradation rate of TC was lower, which was consistent with previous studies. In summary, under the current experimental conditions, CDBPC materials provided excellent photocatalytic degradation effects for TC solutions within 20 mg/L. 7, the R2 values for the pseudo-first-order kinetics ranged from 0.9087 to 0.972, while the R2 values ranged from 0.9438 to 0.9985 for the pseudo-second-order kinetics, suggesting that TC degradation with different pH conditions was more consistent with the pseudosecond-order kinetics. As seen in Table 2, when the initial concentration of TC was 5 mg/L, the degradation rate constant was the largest (2.3 × 10−3 min−1) and the t1/2 time was the lowest (only 87 min). As the initial concentration increased, the degradation rate constant decreased

#### *2.4. Degradation Kinetics of TC* continuously. When the initial concentration of TC was 100 mg/L, the degradation rate

TC degradation was fitted to pseudo-first-order kinetics and pseudo-second-order kinetics models, respectively (Figure 8). The goodness of (R2 ), slope (k), and half-lives (t1/2) of the kinetic equations for the different TC initial concentrations and different initial pH values were calculated in Table 2. For different initial TC concentration conditions, the R2 values of the pseudo-first-order kinetics model ranged from 0.9314 to 0.9937, while that of the pseudo-second-order kinetics model ranged from 0.955 to 0.9949. This suggests that TC degradation was more consistent with the pseudo-second-order kinetics model. As shown in Figure 8c,d, there was a significant difference in the degradation effect under different initial pH conditions. The fitting linear correlation coefficient was only 0.63–0.74 when the pH value was greater than 7. In the range of pH 2–7, the R2 values for the pseudo-first-order kinetics ranged from 0.9087 to 0.972, while the R2 values ranged from 0.9438 to 0.9985 for the pseudo-second-order kinetics, suggesting that TC degradation with different pH conditions was more consistent with the pseudo-second-order kinetics. constant was 0.03 × 10−3 min−1, and t1/2 was 333.3 min. Compared with the initial concentration of TC of 5 mg/L, the degradation rate constant was only 1/76, and t1/2 was 3.83 times longer. As shown in Table 3, a lower pH value was more favorable for the degradation of TC due to the higher OH oxidation potential. When the pH value was 2, the degradation rate was 1.5 × 10−3 min−1, and t1/2 was 67.9 min. When the pH value was 4, the highest degradation rate constant was reached, which was 1.9 × 10−3 min−1, and t1/2 was 53.2 min. As the pH value increased, the degradation rate constant gradually decreased, while t1/2 gradually increased. When the pH value reached 12, the degradation rate constant was 0.05 × 10−3 min−1, and the t1/2 was 1904.2 min. Compared with the pH of 4, the degradation rate constant was 1/38, and the t1/2 was 35.79 times longer.

**Figure 8.** TC degradation of (**a**) pseudo-first-order kinetics and (**b**) pseudo-second-order kinetics with different TC concentration, (**c**) pseudo-first-order kinetics and (**d**) pseudo-second-order kinetics with different initial pH values. **Figure 8.** TC degradation of (**a**) pseudo-first-order kinetics and (**b**) pseudo-second-order kinetics with different TC concentration, (**c**) pseudo-first-order kinetics and (**d**) pseudo-second-order kinetics with different initial pH values.

**Table 2.** TC degradation kinetics parameters with different TC initial concentrations. As seen in Table 2, when the initial concentration of TC was 5 mg/L, the degradation rate constant was the largest (2.3 × 10−3 min−1 ) and the t1/2 time was the lowest (only

**Pseudo-First-Order Kinetic Model Pseudo-Second-Order Kinetic Model K (min−1) R2 t1/2 (min) K (10−3 min−1) R2 t1/2 (min)**

10 0.007 0.9314 99.0 0.9 0.9858 111.1

87 min). As the initial concentration increased, the degradation rate constant decreased continuously. When the initial concentration of TC was 100 mg/L, the degradation rate constant was 0.03 × 10−3 min−1 , and t1/2 was 333.3 min. Compared with the initial concentration of TC of 5 mg/L, the degradation rate constant was only 1/76, and t1/2 was 3.83 times longer.

**Table 2.** TC degradation kinetics parameters with different TC initial concentrations.

As shown in Table 3, a lower pH value was more favorable for the degradation of TC due to the higher OH oxidation potential. When the pH value was 2, the degradation rate was 1.5 × 10−3 min−1 , and t1/2 was 67.9 min. When the pH value was 4, the highest degradation rate constant was reached, which was 1.9 × 10−3 min−1 , and t1/2 was 53.2 min. As the pH value increased, the degradation rate constant gradually decreased, while t1/2 gradually increased. When the pH value reached 12, the degradation rate constant was 0.05 × 10−3 min−1 , and the t1/2 was 1904.2 min. Compared with the pH of 4, the degradation rate constant was 1/38, and the t1/2 was 35.79 times longer.

**Table 3.** TC degradation kinetics parameters with different initial pH.

### *2.5. Free Radical Identification*

BQ, EDTA-2Na and TBA were used as superoxide radical (O2−), hole (h+ ), and hydroxyl radical (OH) scavengers, respectively [41]. As shown in Figure 9, the photocatalytic reaction was almost completely inhibited by the addition of EDTA disodium, indicating that hole (h+ ) was the most dominant active substance for TC degradation. The addition of BQ and TBA also had some effect on the photocatalytic performance, but the effect was not as obvious as that of the hole (h+ ), indicating that ·O2− and OH were also involved in the photocatalytic reaction. The CDs/g-C3N4/BiPO4 composite prepared in this study exhibited an excellent photocatalytic performance.

**Initial Concentration (mg/L)**

20 0.0065 0.9731 106.6 0.5 0.9965 100.0 30 0.0047 0.9485 147.5 0.2 0.9606 166.7 50 0.0023 0.9481 301.4 0.1 0.955 200.0 100 0.0005 0.9937 1386.3 0.03 0.9949 333.3

**Table 3.** TC degradation kinetics parameters with different initial pH.

5 0.0072 0.9459 96.3 2.3 0.9876 87.0 10 0.007 0.9314 99.0 0.9 0.9858 111.1 20 0.0065 0.9731 106.6 0.5 0.9965 100.0 30 0.0047 0.9485 147.5 0.2 0.9606 166.7 50 0.0023 0.9481 301.4 0.1 0.955 200.0 100 0.0005 0.9937 1386.3 0.03 0.9949 333.3

this study exhibited an excellent photocatalytic performance.

 As shown in Figure 9, the photocatalytic reaction was almost completely inhibited by the addition of EDTA disodium, indicating that hole (h+) was the most dominant active substance for TC degradation. The addition of BQ and TBA also had some effect on the photocatalytic performance, but the effect was not as obvious as that of the hole (h+), indicating that ∙O2− and OH were also involved in the photocatalytic reaction. The CDs/g-C3N4/BiPO4 composite prepared in

**Figure 9.** Effect of different quenchers on the photocatalytic degradation of TC. **Figure 9.** Effect of different quenchers on the photocatalytic degradation of TC. *2.6. Proposed Reaction Mechanism*

#### *2.6. Proposed Reaction Mechanism* Figure 10 displays the possible photocatalytic degradation mechanism of T compo-

*2.5. Free Radical Identification*

Figure 10 displays the possible photocatalytic degradation mechanism of T composites. The semiconductor BiPO4 and g-C3N4 were combined to form a heterogeneous structure with CDs attached to its surface and junction. The semiconductor BiPO4 could directly absorb energy with wavelengths less than 320 nm in sunlight, and g-C3N4 could directly absorb energy with wavelengths less than 420 nm. The two materials could absorb solar energy with wavelengths greater than 500 nm by the CDs particles attached to their surfaces to enhance the utilization. After the absorption of solar energy by both materials, electrons in the VB were excited and jumped to the CB, thus generating holes in the VB. The holes (h+ ) could directly degrade TC into intermediate products CO2 and H2O, and the holes (h+ ) could also react with OH− or H2O to produce −OH. Then, the hydroxyl radicals could interact directly with TC degradation [42–44]. The jumped electrons then combined with O2 to produce ·O2−, and superoxide radicals could catalyze TC degradation. sites. The semiconductor BiPO4 and g-C3N4 were combined to form a heterogeneous structure with CDs attached to its surface and junction. The semiconductor BiPO4 could directly absorb energy with wavelengths less than 320 nm in sunlight, and g-C3N4 could directly absorb energy with wavelengths less than 420 nm. The two materials could absorb solar energy with wavelengths greater than 500 nm by the CDs particles attached to their surfaces to enhance the utilization. After the absorption of solar energy by both materials, electrons in the VB were excited and jumped to the CB, thus generating holes in the VB. The holes (h+) could directly degrade TC into intermediate products CO2 and H2O, and the holes (h+) could also react with OH− or H2O to produce −OH. Then, the hydroxyl radicals could interact directly with TC degradation [42–44]. The jumped electrons then combined with O2 to produce ∙O2−, and superoxide radicals could catalyze TC degradation.