Synthesis of a Doped α-Fe₂O₃/g-C₃N₄ Catalyst for High-Efficiency Degradation of Diazinon Contaminant from Liquid Wastes

Abstract

In this work, a hematite/porous graphite carbon-nitride (α-Fe₂O₃/g-C₃N₄) catalyst was synthesized through the doping of hematite loaded onto porous graphite carbon-nitride using a heat treatment process. Then, the ability of catalyst was evaluated to degrade diazinon (DZN) for the first time, mainly via the sonophotocatalytic process. Among the samples, the greatest DZN degradation was observed in the sonophotocatalytic system, which separated 100% of DZN from the aqueous solution after 50 min, while the removal percentages for the sonocatalytic, photocatalytic, and adsorption systems were 72.9, 89.1, and 58.1%, respectively. The results of scavengers showed that both sulfate and hydroxyl radicals (^•OH) participated in removing DZN, although positive holes and negative ^•OH played a major role. Moreover, the removal efficiencies of the target pollutant using the sonophotocatalytic process were higher than those using the photocatalytic, sonocatalytic, and adsorption processes. The reaction profile followed pseudo-first-order kinetics, and the reaction rate coefficient for the sonophotocatalytic system was 2.2 times higher than that of the photocatalytic system and 2.64 times higher than that of the sonocatalytic system. The energy consumption of the sonophotocatalytic system after 60 min was 11.6 kWh/m³, while it was 31.1 kWh/m³ for the photocatalytic system. A DZN removal percentage of 100% was obtained after 50 min under the following conditions: UV intensity of 36 watts, ultrasound frequency of 36 kHz, DZN concentration of 50 mg/L at pH 5, and α-Fe₂O₃/g-C₃N₄ dosage of 0.4 g/L. The catalyst reusability was examined with only a 9.9% reduction in efficiency after eight consecutive cycles. The chemical oxygen demand (COD) and total organic compound (TOC) removal percentages were 95.6% and 88.6%, respectively, and the five-day biochemical oxygen demand (BOD₅)/COD ratio was 0.16 at the beginning of the degradation process and 0.69 at the end of the process. In addition, toxicological experiments showed that degradation of DZN by the sonophotocatalytic process exhibited low toxicity. All results confirmed that the sonophotocatalytic process using α-Fe₂O₃/g-C₃N₄ was a highly efficient process for DZN pollutant removal from liquid wastes.

1. Introduction

Currently, many pollutants are circulating in the environment, including those that have a toxic nature and can result in a significant impact on human health [1]. The spread of these pollutants is mainly due to the increasing pace of growth in industry, agriculture, and the population. From this point of view, many countries are interested in providing efficient treatments to reduce the concentrations of these pollutants to permissible levels, especially in drinking water and aquatic environments [2]. To achieve this, appropriate treatments must be proposed for the wastewater carrying these pollutants to prevent them from spreading to the environment and becoming difficult to contain. Recently, it was found that the amount of toxins produced to control and prevent plant diseases has increased significantly [3]. Widespread use of phosphorus and chlorine toxins, which cause long-term harmful effects on the ecosystem, is one of the most important factors causing the eutrophication of water, the risk to the life of microorganisms, and damage to the respiratory tract and central nervous system of animals [4,5].

DZN, is one of the routinely used insecticides for insect control in the home environment and agriculture [6]. DZN is a nonsystemic toxin that is stable at an approximate pH of 7; its solubility in water is approximately 40 mg/L at ambient temperature, and it is very easy to digest. The present compound of DZN has low volatility in terrestrial and aqueous media and remains stable for several months in these environments [7,8]. From an environmental perspective, DZN is used in large quantities to improve agricultural yield; thus, it may pollute surface- and groundwater. For all these reasons, DZN is categorized as a class II chemical that poses a relative hazard to humans and aquatic organisms, according to the World Health Organization (WHO) [9].

Therefore, advanced methodologies such as photocatalysis and electrooxidation, as well as traditional methods such as adsorption and biodegradation, have been widely utilized for removing these pollutants [10]. Advanced oxidation processes (AOPs) are effective methods for DZN removal [7,8]. Photocatalytic oxidation (PCO) has been used fruitfully for eliminating pollutants, including DZN [8]. Photocatalytic processes are extremely effective methods that use nanoparticles, leading to full mineralization, absolute decomposition of organic pollutants, and CO₂ production under ambient conditions without producing harmful substances [11,12]. Titanium dioxide (TiO₂) is the most frequently used photocatalyst in photocatalytic studies as it is inexpensive, economical, easy to use, and environmentally friendly [13]. However, the use of TiO₂ photocatalysts in photocatalytic studies has disadvantages, such as difficulty in separation, fast recombination of electrons/holes (e⁻/h⁺), and a low bandgap [14,15].

As alternative materials, scientists have worked on graphite-based structures such as graphene (G) and graphene oxide (GO), which have large and hexagonal surfaces and can interact strongly with other molecules [16]. Graphite carbon-nitride (g-C₃N₄) is the most stable carbon nitride allotrope, which is employed in the separation and condensation of organic compounds and thus is suitable for these studies. Graphitic carbon-nitride also has features such as low cost, biaxial and porous structure, environmental compatibility, and chemical stability [17]. In addition to these properties of g-C₃N₄, it can undergo rapid recombination due to the active surface electrons; thus, its recycling and isolation are concerns for the environment. To overcome this problem, the g-C₃N₄ photocatalytic activity was upgraded by doping with semiconductor/metallic/nonmetallic elements [18]. To address this, existing g-C₃N₄ has been combined with metal oxide materials such as TiO₂, Tungsten trioxide (WO₃), Zinc oxide (ZnO), and Molybdenum disulfide (MOS₂) by many researchers to form active photocatalysts in the removal of organic pollutants [19,20]. In addition to their nontoxicity, the advantages of applying semiconductor metals in the improvement of treatment agents include high efficiency, ability to work in different environmental conditions, and high chemical stability over a wide pH range [17,18].

Among these semiconductor materials, hematite (α-Fe₂O₃) has attracted great attention; this is related to characteristics such as low cost, optimum bandgap (2.1 eV), low toxicity, and magnetic separation capabilities. α-Fe₂O₃, which is a decent coupling material for g-C₃N₄, increases the photocatalytic activity under visible light by reducing the photoinduced electron–hole pair recombination rate of the composite after this coupling. According to the literature, there are few studies on the use of α-Fe₂O₃/g-C₃N₄ photocatalysts for degrading dangerous organic pollutants in water [21,22,23].

Several researchers have suggested that the development of a treatment technology that can create a synergistic effect of activating oxidants will meaningfully increase the removal rate of impurities [24,25]. Sonolysis has come to the fore in recent years, in which ultrasound radiation is adopted because of the absence of chemical involvement as a chemical reaction as well as its safety in use [26,27]. During the sonophotocatalytic process, which is a combination of photocatalysis and sonocatalytic processes, it is possible to generate large numbers of ^•OH radicals via ultrasound acoustic cavitation [28]. In addition, the catalytic regions are constantly regenerated by the cavitation process, which means the removal of byproducts and the effects of microcurrent and turbulence. In addition, studies have shown that ultrasound radiation can improve the processes of electron transfer, catalyst dispersion, adsorption, and decomposition of the hydrophobic parts of organic compounds when used in a reactor [29]. Few studies have been found in the literature on the toxicity of samples taken from sonophotocatalytic reactors. Dhanasekar et al. used β-NiMoO₄ in sonophotocatalytic degradation of methylene blue [30]. Ahmad et al. utilized the sonophotocatalysis process to degrade rhodamine B from aqueous solution [31]. Babu et al. informed the removal of organic compounds by a CuO-TiO₂/rGO/US/UV process [32]. Isari et al. evaluated the sonophotocatalytic degradation of antibiotics by WO₃/CNTs composite [33]. Although most of the cited studies reported adequate efficiency of the present process, little information is available for parameters such as mineralization, toxicity assessment, and energy consumption calculations.

In this article, the recovery of the catalyst and the characteristics of the catalyst are completely carried out, which are rarely discussed in studies; in addition, to the toxicity of the effluent after the process is determined. The amount of energy consumption of fog is one of the most imperative items in recent years, and the amount of biological degradability and the amount of mineralization are also calculated herein. Finally, the kinetics of the reactions are also calculated. Therefore, compared to similar studies, this study is comprehensive, and all aspects are considered.

Therefore, the present study aims to use α-Fe₂O₃/g-C₃N₄ to remove diazinon (DZN). In this study, the operating parameters for DZN removal were optimized. The COD, TOC, and BOD₅ concentrations were also determined for the examination of DZN mineralization levels. Additionally, the reusability of the α-Fe₂O₃/g-C₃N₄ photocatalyst was investigated by performing eight sequential reaction cycles. Finally, a scavenger study was carried out to determine the role of radicals in the degradation.

2. Materials and Methods

2.1. Materials and Reagents

Isopropyl alcohol ((CH₃)₂CHOH, >36.0%), ethanol (C₂H₅OH, 99.8%), benzoquinone (C₆H₄O₂, >98%), ethylenediaminetetraacetic acid (EDTA, ≥99.7%), melamine (99%), iron(III) nitrate nonahydrate (Fe(NO₃)₃.9H₂O, 99.99%), sodium chloride (NaCl, ≥99.5%) and sodium sulfate (Na₂SO₄, ≥99%) were provided by Merck, Germany. DZN (≥95%) was supplied by Sigma Aldrich (St. Louis, MO, USA) and used as the target contaminant. Acetonitrile (CH₃CN, 99.8%) and distilled water for DZN analysis by high-performance liquid chromatography (HPLC) were bought from Merck, Darmstadt, Germany.

2.2. Synthesis of α-Fe₂O₃/g-C₃N₄ Composites

The method employed for synthesizing bulk g-C₃N₄ nanosheets has been described in several previous studies [18]. To synthesize α-Fe₂O₃, a 0.1 M Fe(NO₃) solution was made by dissolving Fe(NO₃) (2.5 g) in 50 mL of DI water. NaOH solution was then instilled into the abovementioned solution, and the solution was stirred for 30 min. NaOH has been found to be a robust precipitating agent compared to KOH. After adding NaOH to the precursor solution, good mixing with the precursor solution should be considered, which leads to increasing the solution pH and enhancing the reaction rate. The solution pH was adjusted to a value of 8 after adding NaOH. Before the addition of NaOH, the pH of the solution was 7. At this point, the solution was transferred to a Teflon-lined household microwave oven (2.45 GHz) and illumined for 10 min. Then, centrifugation was used to collect the acquired precipitate. The precipitate was washed with DI water and supreme ethanol separately (a few times) and then dried in a vacuum oven at 45 °C. The final product was a light-colored α-Fe₂O₃ nanopowder. During the synthesis of α-Fe₂O₃/g-C₃N₄ nanocomposites, mixing the exact mass proportion (25, 50, and 100 mg) was considered, and the mixture was crushed through the employment of a mortar and pestle for 3 min. It was then placed in a covered alumina crucible and heated using a suppress heater at 400 °C for 4 h. The considered proportions for g-C₃N₄ and α-Fe₂O₃ were 1:0.5, 1:1, and 1:1.5. The names selected for the ultimate products were F, FG1, FG2, and FG3 (indicating products with 0, 25, 50, and 100 mg of g-C₃N₄, respectively).

The catalyst properties were investigated using XRD (XRD, Rigaku D/Max 2500, Tokyo, Japan) with Cu-Ka radiation (k = 0.15418 nm) in a 2$\mathsf{\theta }$ range of 10–80$°$, SEM/EDX with a Hitachi S4800 microscope (Tokyo, Japan), UV–Vis diffuse reflectance spectroscopy (HACH, DR5000, Loveland, CO, USA), and FTIR spectroscopy with a Thermo Nicolet Avatar 370 FTIR spectrophotometer (Waltham, MA, USA) by means of the KBr method. Determination of the surface area, total pore volume, and pore distribution of α-Fe₂O₃/g-C₃N₄ was fulfilled via N2 adsorption (−196 °C) by employing a TRISTAR-3000 surface area and porosity analyzer (Micromeritics), and a vibrating sample magnetometer (VSM, Lakeshore 665, Westerville, OH, USA) was utilized.

2.3. Catalytic Experiments

All experiments were executed using UV lamps and ultrasound in a one-liter Pyrex reactor containing a 600 mL sample. In each test, a definite amount of α-Fe₂O₃/g-C₃N₄ was added to the sample containing DZN. The solution pH was adjusted using hydrochloric acid and sodium hydroxide (0.1 N). After each test, 10 mL of the sample was taken using a syringe, and after separation of the catalyst using a magnet, it was injected into the HPLC device. The HPLC device was equipped with a C18 column and UV detector (wavelength of 247 nm). The mobile phase contained 60% acetonitrile and 40% deionized water. The injection rate was 1 mL/min. The amount of mineralization in the sonophotocatalytic system was measured by TOC analysis. Additionally, the amount of biodegradability in the system was obtained using BOD₅, COD, and the ratio between them (COD/BOD₅). The amount of compounds from degradation was also obtained by measuring nitrate, sulfate, ammonium, and phosphate ions.

The degradation efficiency (Efficiency %) of DZN was obtained using Equation (1) [34,35].

$$\mathrm{Efficiency} (\%)=\left[\frac{\left({\mathrm{C}}_0-{\mathrm{C}}_{\mathrm{t}}\right)}{{\mathrm{C}}_0}\right]\times 100$$

(1)

C₀ and C_t symbolize the DZN initial concentration and the concentration at a specific degradation time (t, minutes), respectively.

The ability of α-Fe₂O₃/g-C₃N₄ to degrade DZN was probed under various conditions. After defining the optimal values of evaluated parameters, sonolysis, sonocatalysis, photolysis, adsorption, photocatalysis, and sonophotocatalysis processes were employed for assessing DZN removal. Recycling experiments for the sonophotocatalyst were conducted using eight consecutive reaction cycles. After each application, catalysts were collected using a magnet. They were then rinsed with ethanol/distilled water (50/50 V/V) to remove impurities. Finally, for reuse, the drying process in an oven was applied at 60 °C for 2 h. Trapping experiments with isopropyl alcohol (IPA), Ethylenediaminetetraacetic acid (EDTA), and p-benzoquinone (BQ) scavengers were performed to categorize the species active in the removal of DZN. Toxicity evaluation of the treated solution was performed by microbial culture of E. coli. In an autoclave, several lactose broth tubes were sterilized. Incubation of E. coli in Eosin Methylene Blue (EMB) agar medium was done at 37 °C for 24 h. To appraise the toxicity, some tubes were used for the toxicity tests (10 mL of culture medium + 1 mL of reactor output + one microbial loop), and some were used for the control tests (10 mL of culture medium + one microbial loop). After incubation at 37 °C, measuring the absorption of the tubes was fulfilled by a UV–Vis spectrophotometer at 600 nm.

3. Results and Discussion

3.1. Characterization Results

The optical absorption and bandgap of the catalysts were scrutinized through the employment of UV-DRS, and the reflectance diagram is illustrated in Figure 1a. It is clear that all specimens had a stronger absorption in the visible region. The absorption edges for α-Fe₂O₃ and α-Fe₂O₃/g-C₃N₄ were found at 456 and 461 nm, respectively [18]. A shift was observed toward the longer wavelength side. The remarkable redshift detected within the absorption edge is indicative of a reduction in the bandgap of the catalyst. Based on the UV diagram (1b), the bandgap was estimated using the Kubelka-Munk (K-M) relation, and the bandgap was determined to be 2.72, 2.7, 2.53, 2.41, and 2.13 eV for g-C₃N₄, F, FG1, FG2, and FG3, respectively. The smaller bandgap of the FG3 catalyst is due to the combination of the high amount of g-C₃N₄ with α-Fe₂O₃, which induces an imaginable charge transfer between g-C₃N₄ and Fe₂O₃. Since FG3 has a lower bandgap, the study is performed using FG3.

Figure 1c delineates the N₂ adsorption–desorption analysis and the pore size distribution of α-Fe₂O₃ and α-Fe₂O₃/g-C₃N₄ catalysts. It is observed that both samples show type IV isotherms, which indicates the mesoporous nature of the sample [21]. The surface area of α-Fe₂O₃, g-C₃N₄, and α-Fe₂O₃/g-C₃N₄ catalysts are 21.2, 124, and 107 m²/g, respectively. In addition, the pore size of the α-Fe₂O₃/g-C₃N₄ catalyst is 26 nm. A high surface area of the α-Fe₂O₃/g-C₃N₄ catalyst is desirable to enhance the adsorption reaction.

The morphology and nanostructure of the samples were analyzed by SEM, TEM, and HRTEM, and the SEM results are represented in Figure 1d. As publicized by the SEM, α-Fe₂O₃ was regular and spherical, while the synthesized α-Fe₂O₃/g-C₃N₄ nanocomposite was irregular and had a rugged surface. This could be due to the formation of α-Fe₂O₃ nanoparticles on the g-C₃N₄ catalyst surface [18]. The TEM image of the composite displays the thriving dispersion of α-Fe₂O₃ nanoparticles on the surface of g-C₃N₄ nanosheets, with an average size of 20–30 nm (Figure 1e). The α-Fe₂O₃ particle coating can reduce the aggregation of g-C₃N₄ bulk sheets. The HRTEM image shows that the interfacial distance is approximately 0.36 nm (Figure 1f), which corresponds to the (012) plane of α-Fe₂O₃. The HRTEM image reveals a close relationship between g-C₃N₄ and α-Fe₂O₃, which indicates the formation of the α-Fe₂O₃/g-C₃N₄ composite. Intimate contact between g-C₃N₄ and α-Fe₂O₃ is useful for separating and transporting photogenerated charge carriers [17]. The HRTEM and TEM images confirm the SEM results and show that α-Fe₂O₃ particles are uniformly dispersed on the g-C₃N₄ surface. In addition, the EDX image of the catalyst is publicized in Figure 2a. Based on the mentioned figure, the catalyst is composed of carbon, nitrogen, oxygen, and iron elements.

In Figure 2b, FT-IR spectra related to g-C₃N₄ and α-Fe₂O₃/g-C₃N₄ nanocomposites are shown. Similar chemical functional groups were observed, which can be explained by the lack of obvious differences between the two spectra. A wide absorption band is observed at 3446 cm⁻¹; it originates from the O-H stretching vibration of the water molecules adsorbed on the surface of the nanocomposites. An absorption peak was detected at approximately 2900 cm⁻¹, which corresponds to the stretching vibration of C-H [17]. At approximately 1600 cm⁻¹, the characteristic peak of g-C₃N₄ was observed, which is ascribed to the stretching vibration of C-N heterocycles. In Fe₂O₃-doped g-C₃N₄, the mentioned peak was weaker, which is indicative of the alignment of iron ions with C-N aromatic rings. Another peak observed at 1050 cm⁻¹ is related to the C-O stretching vibration. At approximately 670 cm⁻¹ in the spectrum of the α-Fe₂O₃/g-C₃N₄ nanocomposite, there was a weak peak ascribed to the Fe-O stretching vibration state [21]. The results confirmed the formation of binary α-Fe₂O₃/g-C₃N₄ nanocomposites due to the successful interaction of iron oxide nanoparticles with pure g-C₃N₄.

Magnetometer vibrating sample (VSM) analysis was performed to evaluate the magnetic properties of the α-Fe₂O₃ and α-Fe₂O₃/g-C₃N₄ catalysts, and the results are presented in Figure 2c. The results of the above analysis showed that the maximum saturation magnet values for α-Fe₂O₃ and α-Fe₂O₃/g-C₃N₄ were 61 and 21.5 g/emu, respectively. Therefore, it can be acknowledged that the catalyst synthesized in the present study has very good magnetic properties for magnetic separation. Based on the reduction in the obtained α-Fe₂O₃ saturation magnetic curve, it is confirmed that α-Fe₂O₃ is located on the surface of the α-Fe₂O₃/g-C₃N₄ catalyst.

The phase structures of α-Fe₂O₃, α-Fe₂O₃/g-C₃N₄, and g-C₃N₄ were assessed by XRD, and the patterns are displayed in Figure 2d,e. The XRD spectrum corresponding to g-C₃N₄ represents two peaks at 27.7° and 13.1°, which belong to the graphitic C₃N₄ hexagonal phase (JCPDS 87-1526). The peak detected at 27.7°, which is related to the g-C₃N₄ (002) plane, is explained by the interlayer stacking of the conjugated aromatic system. The appearance of the peak at 13.1° is related to the g-C₃N₄ (100) plane, which is due to an in-plane structural packing motif. In the XRD spectrum of α-Fe₂O₃, the peaks observed at 24.18°, 33.15°, 35.75°, 40.93°, 49.43°, 54.02°, 57.56°, 62.51°, and 64.05° were related to the (012), (104), (110), (113), (024), (116), (018), (214), and (300) planes of hexagonal α-Fe₂O₃ (JCPDS No. 33-0664), respectively. Peaks of Fe₂O₃ and g-C₃N₄ were also seen in the synthesized nanocomposites, indicating that Fe₂O₃ was doped on g-C₃N₄.

Through XPS, the combination of g-C₃N₄ and α-Fe₂O₃ in the α-Fe₂O₃/g-C₃N₄ composite was inspected. Fe, O, N, and C are elements detected in the XPS spectrum of α-Fe₂O₃/g-C₃N₄ (Figure 3a). The binding energies of C 1s were detected at 284.1, 288.1, and 288.9 eV. They were related to sp2 hybridized C atoms in C-C, C=N, and C=O, respectively (Figure 4b); defect-containing sp² hybridized carbon atoms present in the graphitic domains and pure graphitic sites in the CN matrix may be the reason for those binding energies. The binding energies of N 1s were 397.9 eV (assigned to sp2 hybridized N atoms in C=N), 398.9 eV (related to sp³ hybridized N atoms in C-[N]3), and 400.6 eV (assigned to sp3 hybridized N atoms in C-NHX) (Figure 4c). In Figure 4d, the high-resolution spectrum of O 1s is presented. The binding energies were 531.4 and 527.9 eV, which were related to sp2 hybridized O atoms in C=O and latticed oxygen in α-Fe₂O₃. Moreover, the binding energy of 529.7 eV was observed for lattice oxygen in pure α-Fe₂O₃ (Figure 4e); however, for the binding energy of lattice oxygen in α-Fe₂O₃/g-C₃N₄, a negative shift of 1.8 eV, compared to that of pure α-Fe₂O₃, was detected. As depicted in Figure 4f, the peaks of Fe 2p are at 709.8 and 724.1 eV, which can be found to be Fe 2p_3/2 and Fe 2p_1/2, respectively. According to Figure 4g, the binding energies of Fe2p in pure α-Fe₂O₃ were 710.5 and 725.6 eV [21]. The shift that occurred to low binding energy for Fe 2p is possibly related to the formation of Fe-N. This results in the enhancement of the electron density on the Fe, which is explained by the fact that N has a relatively small electronegativity compared to O; this indicates the possibility of forming chemical bonds between two components instead of a physical mixture [17].

3.2. Study of Effects of Operating Parameters

3.2.1. Ultrasonic Wave Frequency

In this study, several experiments were conducted to examine the effect of different ultrasonic wave intensities (20 to 50 kHz) (Figure 4a) (conditions: initial concentration of DZN of 100 mg/L, UV light intensity of 36 W, α-Fe₂O₃/g-C₃N₄ mass of 0.4 g/L, and solution pH of 5). The degradation efficiency of the 100 mg/L DZN solution increased from 89.7% at 20 kHz to 100% at 50 kHz after 60 min of radiation time. This trend is similar to the work by Liu et al. (i.e., enhancing the percentage of degradation exhibited a direct link to the increase in ultrasonic frequency) [36]. Lim et al. inspected the removal of chlorinated compounds and concluded that the degradation efficiency increased as a result of increasing the frequency from 35 to 300 kHz due to high reactive oxygen species [37]. An increase in the DZN removal by increasing the ultrasonic wave intensity in the obtained results is due to the following reasons: (a) the number of cavitation bubbles is increased [38], (b) the bubbles collapse due to the increase in power, resulting in a large number of ^•H and ^•OH radicals produced [39], and (c) the catalyst surface area increases as a result of de-aggregation of α-Fe₂O₃/g-C₃N₄ caused by ultrasonic irradiation [21].

3.2.2. Ultraviolet Light Intensity

Another parameter that is effective in photocatalytic processes is the intensity of ultraviolet light. To examine the effect of ultraviolet light intensity on the system efficiency, experiments were carried out with UV lamps with different powers between 8–36 W (conditions: DZN amount of 100 mg/L, α-Fe₂O₃/g-C₃N₄ mass of 0.4 g/L, solution pH of 5, and ultrasonic waves of 35 kHz). The effect of UV light intensity on DZN degradation under optimum conditions was investigated and is shown in Figure 4b. This figure indicates that when the irradiance is increased from 8 to 36 W within 60 min of irradiation, there is a corresponding clear increase in efficiency from 83.9 to 100%. Electrons are more excited with increasing lamp power. It enhances the number of active ^•OH, which are the main reason for the degradation efficiency of organic pollutants in the photocatalytic treatment systems [40]. The results of the study by Nasirian et al., which investigated the effect of light intensity on the degradation of a dye using the Fe₃O₄/TiO₂/UV catalyst, show a similar trend [41].

3.2.3. Catalyst Dosage

It is stated in most reports that another parameter that has an effect on degradation efficiency in advanced oxidation processes (AOPs) is catalyst dosage [14]. Therefore, in this study, the influence of catalyst dosage changes on the diazinon (DZN) degradation efficiency with a reaction time of 0 to 60 min was evaluated (Figure 4c). This experiment was performed considering constant experimental conditions: UV light intensity of 24 W, DZN amount of 100 mg/L, solution pH of 5, and ultrasonic waves of 35 kHz. The obtained results showed that the degradation efficiency amplified with the increase in the catalyst dose. The detected phenomenon can be explicated by the continuous increase in the formation of active ^•OH [42]. In addition, Figure 1c indicates that the optimal catalyst dose is 0.4 g/L. The reactor volume and the amount of aqueous solution are effective in determining the optimal dose loading of a catalyst. It is expected that the same optimal α-Fe₂O₃/g-C₃N₄ dosage will be valid for all parameters, since the same reactor is used for all parameters examined in DZN removal. In addition, the DZN degradation efficiency using α-Fe₂O₃/g-C₃N₄ at optimum conditions was compared with the degradation efficiency using only Fe₂O₃ (0.4 g/L). The results showed that α-Fe₂O₃/g-C₃N₄ has a higher ability to degrade DZN molecules than Fe₂O₃ particles, in which the ability of these particles only reached 18.89%.

3.2.4. Initial Diazinon Concentration

The initial concentration of the pollutant is one of the parameters that affect the efficiency of the photocatalytic working system. In this study, the effect of initial diazinon (DZN) concentration was probed under constant operating conditions (α-Fe₂O₃/g-C₃N₄ dose of 0.4 g/L, solution pH of 5, UV light intensity of 15 W, and ultrasonic waves of 35 kHz). Figure 4d shows the effect of a change in the DZN initial concentration on the photocatalytic degradation efficiency. In this study, DZN solutions with concentrations of 25, 50, 75, and 100 mg/L were put into the reactors, and the degradation efficiency of the processes was investigated. According to the results obtained in the presence of a constant amount of catalyst, since the increase in DZN concentration caused a decrease in the reaction rate, the rate of DZN decomposition decreased. For example, while DZN solution with a 25 mg/L concentration was removed with 100% efficiency after 50 min, DZN solution with an initial concentration of 100 mg/L could be eliminated with 87% efficiency after the same time. The result was possibly due to pollutant adsorption on the catalyst surface [43]. During adsorption, the number of photons present on the catalyst surface decreases. The decrease in the number of photons leads to a decrease in h⁺ pairs, which reduces the formation of OH species, which play an active role in the degradation of DZN, thus reducing DZN removal [7].

In the advanced oxidation process (AOP), another parameter that affects the impurity removal efficiency is the reaction time. To examine the effect of time on the efficiency of DZN removal, studies were carried out for 60 min under the optimum conditions obtained in previous experiments. According to the results (Figure 4d), the elimination efficiency is improved with increasing reaction time. The reaction rate was high in the first 20 min of the process, and DZN solution with a 100 mg/L concentration reached a 90.25% removal efficiency at the end of this period. The reason why the reaction rate decreases later on is the short lifetime of ^•OH radicals, which can only oxidize organic molecules at or near the point of generation [44]. The presence of more impurities at the beginning of the photocatalytic processes will increase the oxidation of ^•OH radicals, and the yield changes per unit time in the ^•OH radical formation region will decrease with the decrease in the impurity in the later process. At the end of 50 min, all (100%) of the 25 mg/L DZN solution was removed due to sufficient SO₄^•− and ^•OH production. Due to the technique of oxidation–reduction reactions, it is a commonly established fact that ^•OH production ends and SO₄^•− continues; the free radicals are formed, and pollutants are decomposed [7]. Lin et al. [45] studied the effect of contact time in the ultrasonically activated PS process for phenol removal. According to their results, the percentage of removal increased from 35% to 88% when the mixing time was increased from 10 to 90 min.

3.2.5. Comparative Study of Diazinon Degradation under Different Processes

To survey the effect of different processes on diazinon (DZN) degradation, six different oxidation processes were investigated under optimal conditions from the previous steps, and the results are displayed in Figure 4e. It shows that DZN removal after 60 min was 27.4%, 49.7%, 57.9%, 72.9%, and 95.5% using photons, catalysts, US, US + CATA, and PHOTO + CATA, respectively. However, the removal of 100 mg/L DZN solution resulted in 100% efficiency due to the simultaneous formation of SO₄^•− and ^•OH radicals using the US + PHOTO + CATA hybrid process. As a result, the concurrent use of US, PHOTO, and CATA is very important in terms of energy consumption, as the combination of mentioned methods will reduce the time to attain the highest removal efficiency.

In order to better evaluate the performance of the sonophotocatalytic process based on α-Fe₂O₃/g-C₃N₄, the comparison of the efficiency of the processes in removing different organic pollutants is shown in

Table 1. The degradation of organic compounds by AOPs based on α-Fe₂O₃/g-C₃N₄.

Pollutant	Catalyst Dose (g/L)	Time (min)	Method	Degradation (%)	Ref.
Bisphenol A (0.5 mM)	0.5	60	Persulfate activation	92.2	[23]
Direct red 81 (0.05 mM)	1.25	120	Photocatalytic	≈80	[46]
Rhodamine B (10 mg/L)	1	120	Photocatalytic	90	[47]
Orange II (50 mg/L)	-	60	Photo-Fenton	≈80	[48]
DZN (100 mg/L)	0.4	60	Sonophotocatalytic	100	Present study

. As can be seen, the present study is confirmed by similar studies that α-Fe₂O₃/g-C₃N₄ can be considered as a suitable catalyst for the degradation of pollutants by different oxidation processes. It should be emphasized that the chemical structure of DZN is complex and stable compared to other pollutants in Table 1. Therefore, it is difficult to remove and mineralize it through different processes. In addition, the amount of DZN in our study was 100 mg/L, which was degraded more than 100% in the reaction time of 60 min in the presence of a 0.4 g/L catalyst. Based on the mentioned cases, it can be reported that α-Fe₂O₃/g-C₃N₄ photocatalyst is a promising composite in the degradation of DZN from aqueous solution.

3.2.6. Solution pH

One of the factors affecting the removal of contaminants is the initial pH of the solution. In AOPs, solution pH is extremely effective in the formation of SO₄^•− and ^•OH, the stability of oxidants, and the determination of the state and type of iron in the solution [15]. To determine the effect of pH on DZN removal, many experiments were carried out with different values between pH 3 and 11. The results are shown in Figure 4f (conditions: α-Fe₂O₃/g-C₃N₄ dose of 0.4 g/L, initial concentration of DZN of 100 mg/L, ultrasonic waves of 35 kHz, and UV light intensity of 36 W). In the photocatalytic and sonophotocatalytic systems studied for DZN removal, it was concluded that the alkalinity of the aqueous solution decreases the reaction rate, and higher removal occurs under acidic conditions [7]. Charge changes on the catalyst surface naturally affect catalysis, which is a surface phenomenon. In this study, the pH_PZC of the catalyst was determined to be 6.7. Consequently, the surface charge of the catalyst is positive when the pH is less than 7 and negative when it is greater than 7. Therefore, the catalyst strongly adsorbs DZN molecules under acidic (pH 3–5) conditions due to the positively charged electrostatic attraction [49]. In the reaction between DZN adsorption molecules and sonophotocatalytic reactive radical species formed on the α-Fe₂O₃/g-C₃N₄ surface, DZN and the catalyst act as a strong Lewis base and a strong Lewis acid, respectively. With decreasing pH, weakened azo bonds are easily broken, and with increasing pH, the rate of degradation decreases due to the negative charges formed between DZN and the catalyst [50].

3.3. Investigation of the Mineralization Behavior of Diazinon

Wastewater treatment has many purposes, the main one of which is to completely decompose pollutants and convert them into biodegradable/safe products. In this work, the analysis of treated DZN-containing solutions in terms of toxicity and mineralization was carried out by a series of test runs. TOC analysis was used to assess the mineralization levels of DZN in the sonophotocatalytic process, and the results obtained are presented in Figure 5a. The graph in Figure 1g shows the degradation and mineralization process of DZN after 120 min under optimal conditions, showing that 100% mineralization did not occur and intermediate products of DZN were formed. The results shown in Figure 1g demonstrate that when the time was increased from 20 min to 120 min, the COD and TOC removal efficiencies were 95.6% and 88.6%, respectively, over time. The fact that the reduction efficiency of TOC is less than the reduction efficiency of DZN is related to the reduction of DZN to less toxic products containing TOC [51]. The TOC and COD concentrations, which were at initial concentrations of 61.9 and 351.2 mg/L, decreased to 7.07 and 15.4 mg/L, respectively, after 120 min. This event indicates that DZN and its intermediates are eventually converted to CO₂ and H₂O in the process studied. As a result, the sonophotocatalysis process is thought to be an operative method for eliminating DZN and its byproducts from aqueous solution, as well as providing appropriate mineralization.

3.4. Investigation of the Degradation Behavior of Diazinon

The toxicity of wastewater comes first among the parameters that make it difficult to work in biological processes. Industrial waste is not biodegradable, as observed by most scientists. Therefore, industrial wastewater is subjected to advanced treatment processes before being converted into biodegradable wastes [7]. The BOD₅/COD ratio is used to determine this parameter. Wastewater with a BOD₅/COD ratio above 0.4 is considered to be biodegradable [30]. A large number of test runs were carried out under optimum conditions, and the biodegradation behavior of the system was analyzed. Figure 5a shows the experimental results. The biodegradability (BOD₅/COD ratio) of the DZN solution with time showed changes, as shown in Figure 5a. As the reaction time increased from 0 to 120 min, the BOD₅/COD ratio increased from approximately 0.16 to 0.69. Therefore, this improvement in the BOD₅/COD ratio showed that the starting solution is biodegradable as a result of the US + PHOTO + CATA system. Based on the BOD₅/COD results, it can be concluded that the suggested degradation system (sonophotocatalytic process) was efficient in terms of converting the nonbiodegradable DZN molecules into biodegradable byproducts. Therefore, this treatment system can be installed as a pretreatment process before the secondary treatment of industrial or municipal wastewater prior to discharge into the surrounding environment.

In a study by Dbira et al. [52] for tannic acid removal by UV/S₂O₈²⁻ and US/H₂O₂/Fe²⁺, ratios of 0.48 and 0.56 were obtained in effluent UV/S₂O₈²⁻ and UV/H₂O₂/Fe²⁺ systems with a BOD₅/COD initial ratio of 0.2, respectively, and it was concluded that these systems are suitable for the biodegradation of DZN solution.

3.5. The Stability and Reusability of Fe₂O₃/g-C₃N₄

Being stable against metal ion leaching under operating conditions has been known as one of the requirements for employing heterogeneous catalysts in wastewater treatment. Lack of proper stability and continuous and gradual leaching might be associated with the deactivation of the catalyst and further water pollution. In addition to total leaching, the reusability of the catalyst is a critical feature for a heterogeneous catalyst [53,54]. Therefore for checking the reusability, the catalyst was withdrawn from the reaction mixture between each cycle by filtration, washed several times and dried as detailed in the experimental part, and then reused. The results of the studies evaluating the reusability of the synthesized α-Fe₂O₃/g-C₃N₄ are presented in Figure 5b. Experiments were performed under optimal conditions at a concentration of 100 mg/L. As verified by the results obtained, there was a gradual decrease in the catalytic activity of α-Fe₂O₃/g-C₃N₄; therefore, at the end of the eighth cycle, the removal efficiency decreased by only 9.87%. This 9.87% decrease indicates that the synthesized photocatalyst had both good stability and recyclability. Furthermore, after the eighth cycle, it is observed that the iron leaching from the catalysts is less than 10–20 mg, which is a negligible value. The small amount of g-C₃N₄ loss caused the dissociation of the catalytic activator sites, resulting in a loss of catalytic activity [55]. The catalytic activity of the catalyst can be affected by the adsorption of the intermediates, which is due to the fact that its surface can be covered by DZN and the degradation intermediates, resulting in inhibition of the reaction, and the catalyst is used without washing in every study [7].

3.6. Scavenger Test

The most active species generated during the DZN degradation were investigated by scavenging tests in the analysis of the degradation mechanism. Experiments were performed under optimal conditions at a concentration of 100 mg/L. Detection of active species responsible for decolorizing DZN was performed through the employment of capture reagents, e.g., BQ, EDTA, and IPA; the agents mentioned were used for superoxide (^•O₂⁻) radicals, holes (h⁺), and ^•OH, respectively [16]. ^•OH radicals were the most effective determinants of color removal of the DZN solution, as seen in Figure 5c. Adding isopropyl alcohol, an ^•OH scavenger, to the DZN solution resulted in a reduction in the decolorization rate. However, when other scavengers are added, a decrease in degradation is observed. Therefore, although ^•OH is the most effective species among the active species, combining these active species might further progress the degradation efficiency [7].

3.7. Determining Energy Consumption

One of the most important factors for evaluating the proficiency of treatment systems is the cost of the system. In recent years, some of the important problems related to treatment methods and comparison of different methods have been energy consumption and costs related to wastewater treatment (per cubic meter of treated wastewater).

The energy consumption (EC) in kilowatt-hours per cubic meter during sonophotocatalytic and photocatalytic processes was estimated through applying the following equation [56,57]:

$$\mathrm{EC} =\frac{\mathrm{P} \times \mathrm{t}}{\mathrm{V} \times \log (\frac{{\mathrm{C}}_0}{{\mathrm{C}}_{\mathrm{e}}})}\times \left(\frac{1000}{60}\right)$$

(2)

where V refers to the volume of DZN solution in the photochemical reactor, t represents the time taken to treat the solution, and P represents the input power (kW) of the treatment system.

The amount of energy consumed for both systems is shown in Figure 5d. The amount of energy consumed in 60 min for the photocatalytic system to remove 100% of the 50 mg/L DZN solution was 11.6 kWh/m³, while at the same time and at the same concentration, it was equal to 31.1 kWh/m³ for the photocatalytic system, and the results show that the energy consumption in the sonophotocatalytic system is lower than that in the photocatalytic system.

In the study piloted by Lin et al. [58] for eliminating phenol by a TiO₂ catalyst, the amount of energy consumption was detected to be 145.5 kWh/m³, and 80% removal of phenol was observed; the results showed that the amount of energy consumption increases with increasing time. In another study, Kumar et al. utilized ZnO nanoparticles to remove acid black 1 dye, and the energy consumption was equal to 121.2 kWh/m³ [59]. Yilmaz et al. [10] appraised the removal of acid blue 113 dye and reported that the amount of energy consumption during treatment was calculated between 10 and 60 min; their results were indicative of the fact that with increasing time, the amount of energy consumption increased from 21.2 to 43.4 kWh/m³. In addition, in the study of Arghavan et al. conducted for the elimination of tamoxifen with FeNi₃@SiO₂@ZnO, the amount of energy consumption was calculated at contact between 30 and 120 min; based on their reports, with increasing time, the amount of energy consumption increased from 43.1 to 96.7 kWh/m³ [60].

3.8. Kinetic Degradation of Diazinon

The chemical reaction kinetics of organic compounds can be expressed by the pseudo-first-order kinetics (PFO) model or the Hinshelwood-Langmuir model. This model is related to the rate of degradation and concentration of organic compounds, which is expressed as Equation (3) [29].

$$\mathrm{r}=\frac{{\mathrm{d}}_{\mathrm{C}}}{{\mathrm{d}}_{\mathrm{t}}}=\frac{{\mathrm{K}}_{\mathrm{d}}{\mathrm{k}}_{\mathrm{r}}\mathrm{C}}{1+\mathrm{KC}}$$

(3)

where k_r is the original constant rate (1/min) and K_d is the adsorption equilibrium constant (l/mg).

When there is relatively low adsorption or low concentration of organic compounds, Equation (3) can be simplified or expressed as first-order kinetics at a known constant rate (K, 1/min) (Equation (4)).

$$\ln \frac{{\mathrm{C}}_0}{{\mathrm{C}}_{\mathrm{t}}}=-\mathrm{Kt}$$

(4)

Based on the PFO model, as shown in Figure 5e, the ln C₀/C_e data are drawn against the reaction time as a straight line, while the slope of the diagram shows the rate constant (k). Accordingly, the degradation of DZN by different processes according to the high regression coefficient is performed in accordance with PFO kinetics. In addition, the reaction rate constant for the sonophotocatalytic system was 0.66, and it was 0.3 and 0.25 1/min for photocatalytic and sonocatalytic systems, respectively; as seen, the rate of DZN degradation by the sonophotocatalytic system is more than double that of other systems. In a study conducted by Al-Musawi et al. for degrading acid blue 113 dye, the results showed that the sonophotocatalytic rate was approximately three to four times higher than that of photocatalytic and sonocatalytic systems [28].

3.9. Toxicological Experiments Using Algal Growth

To evaluate the toxicity of products produced from the degradation of DZN by various sonophotocatalytic, photocatalytic, and sonocatalytic systems, algal growth testing was performed in DZN solution (after one hour of DZN degradation by various processes). Additionally, two solutions were considered, including a solution without DZN and a solution with DZN (50 mg/L). Experiments were performed using the alga Microcystis aeruginosa according to the Organization for Economic Co-operation and Development (OECD) guideline [61]. The test conditions were a light intensity of 8000 lux, a temperature of 20 °C, and an initial alga concentration of 1.1 $\times$ 10⁵ cells/mL. The results of algal growth are shown in Figure 5f; except for the sonophotocatalytic process, which exhibited algal growth similar to the DZN-free control solution, there were still toxic compounds in the effluent of other processes (despite reducing the toxicity of the effluent) that inhibited algal growth compared to the control solution. The algae cell density in the control sample increased from 1.1 $\times$ 10⁵ to 98.1 $\times$ 10⁵ cells/mL, and in the sample of the sonophotocatalytic process, this value was equal to 97.8 $\times$ 10⁵ cells/mL, which indicates a significant reduction in toxicity in the effluent. In the untreated DZN sample, the algae cell density decreased after eight days and reached 1.3 $\times$ 10². However, in the samples treated with sonocatalytic and photocatalytic processes, the algae cell density reached 65.4 $\times$ 10⁵ and 74.3 $\times$ 10⁵ cells/mL, respectively, which indicates a significant reduction in toxicity, but the reduction in toxicity was less than that of sonophotocatalytic processes. In a study conducted by Sajjadi et al. for removing DZN by a Fe₃O₄@MOF-2 catalyst, there was a significant reduction in toxicity; however, the toxicity of DZN was not completely eliminated by the photocatalytic process [2].

3.10. Proposed Degradation Mechanism

In order to propose a possible mechanism, the conduction band (E_CB) and valence band (E_VB) values of α-Fe₂O₃/g-C₃N₄ particles should be determined. The positions of E_CB and E_VB were determined according to the equations described by Mulliken electronegativity theory (Equations (5) and (6)).

$${\mathrm{E}}_{\mathrm{CB}}= \mathsf{\chi }-{ \mathrm{E}}_{\mathrm{C}}-0.5{\mathrm{E}}_{\mathrm{g}}$$

(5)

$${\mathrm{E}}_{\mathrm{VB}}={\mathrm{E}}_{\mathrm{CB}}+{ \mathrm{E}}_{\mathrm{g}}$$

(6)

where E_CB and E_VB represent the potential of CB and VB particles, χ indicates the electronegativity of the catalyst, E_c demonstrates the energy of free electrons (4.5 eV) in the hydrogen scale, and Eg shows the band gap of the photocatalyst.

According to previous studies [46,62], χ values for Fe₂O₃ and g-C₃N₄ were reported as 5.89 and 4.73 eV, respectively. According to Eg of Fe₂O₃ (2.7 eV) and g-C₃N₄ (2.72 eV) obtained by DRS analysis, the E_CB and E_VB values were computed, according to the above equations, to be +0.04 and +2.74 eV for Fe₂O₃ and −1.13 and +1.59 for g-C₃N₄, respectively.

Considering our reported results, potential calculations, and earlier studies [7], a possible DZN sonophotocatalytic mechanism is presented in Figure 6. According to this Figure and the suitable surface area of the catalyst, a part of the DZN molecule can be adsorbed on the α-Fe₂O₃/g-C₃N₄ surface. At the same time, under UV light radiation, both Fe₂O₃ and g-C₃N₄ particles can be excited and participate in producing electron-hole pairs. Since VB Fe₂O₃ is more positive than VB g-C₃N₄ and the conduction band of g-C₃N₄ is more negative than Fe₂O₃, the electron in CB g-C₃N₄ is transferred to CB Fe₂O₃, while the hole is transferred to CB g-C₃N₄ (Equation (7)). This transfer process reduces the electron-hole recombination and advances the charge separation efficiency. In addition, the photogenerated electron on the hematite surface reacts with dissolved oxygen (O₂) to produce O₂^•− (Equation (8)) [63]. On the other hand, the holes produced on the g-C₃N₄ surface can react with H₂O and OH^- molecules for producing ^•OH (Equation (9)) [28].

In the proposed mechanism, ultrasonic radiation causes the water molecules in the sonophotocatalytic reactor to decompose and release ^•OH and H₂O₂ to the environment (Equations (10)–(13)) [64,65]. Meanwhile, the energy required to increase the formation of e−/h+ pairs is provided by catalyst-coupled ultrasonics (Equation (14)). During these steps, UV light degrades H₂O₂ and produces ^•OH [66]. Finally, the oxidation of DZN molecules adsorbed on the α-Fe₂O₃/g-C₃N₄ active sites is carried out by the active species (H, O₂^•− and h+) (Equation (15)).

$$\mathsf{\alpha }-{\mathrm{Fe}}_2{\mathrm{O}}_3/\mathrm{g}-{\mathrm{C}}_3{\mathrm{N}}_4+\mathrm{hv}\to \mathsf{\alpha }-{\mathrm{Fe}}_2{\mathrm{O}}_3\left({\mathrm{e}}_{\mathrm{C}\mathrm{B}}^{-}\right)+ \mathrm{g}-{\mathrm{C}}_3{\mathrm{N}}_4\left({\mathrm{h}}^{+}\right)$$

(7)

$$\mathsf{\alpha }-{\mathrm{Fe}}_2{\mathrm{O}}_3\left({\mathrm{e}}^{-}\right)+{\mathrm{O}}_2\to {\mathrm{O}}_2^{-·}+\mathsf{\alpha }-{\mathrm{Fe}}_2{\mathrm{O}}_3$$

(8)

$$\mathrm{g}-{\mathrm{C}}_3{\mathrm{N}}_4({\mathrm{h}}^{+})+{\mathrm{H}}_2\mathrm{O}/{\mathrm{OH}}^{-}\to {}^{•}\mathrm{OH}+\mathrm{g}-{\mathrm{C}}_3{\mathrm{N}}_4$$

(9)

$$\mathrm{ultrasound})))+{\mathrm{H}}_2\mathrm{O}\to {}^{•}\mathrm{OH}+{\mathrm{H}}^{•}$$

(10)

$${\mathrm{H}}^{•}+{\mathrm{O}}_2\to {\mathrm{HO}}_2^{•}$$

(11)

$${\mathrm{HO}}_2^{•}+{\mathrm{HO}}_2^{•} \to { \mathrm{H}}_2{\mathrm{O}}_2+{\mathrm{O}}_2$$

(12)

$${\mathrm{H}}_2{\mathrm{O}}_2+\mathrm{hv} \to 2{}^{•}\mathrm{OH}$$

(13)

$$\mathsf{\alpha }-{\mathrm{Fe}}_2{\mathrm{O}}_3/\mathrm{g}-{\mathrm{C}}_3{\mathrm{N}}_4+\mathrm{ultrasound})))\to \mathsf{\alpha }-{\mathrm{Fe}}_2{\mathrm{O}}_3/\mathrm{g}-{\mathrm{C}}_3{\mathrm{N}}_4\left({\mathrm{e}}^{-}+{\mathrm{h}}^{+}\right)$$

(14)

$${}^{•}\mathrm{OH}/{\mathrm{h}}^{+}/{{\mathrm{O}}_2}^{•-}+\mathrm{DZN}\to {\mathrm{CO}}_2+{\mathrm{H}}_2\mathrm{O}+\mathrm{degradable} \mathrm{byproducts}$$

(15)

4. Conclusions

The degradation of diazinon (DZN) was performed using the α-Fe₂O₃/g-C₃N₄ catalyst. The influence of reaction time, initial concentration of DZN, pH, radiation intensity, frequency intensity, and catalyst mass was investigated, and the catalyst characteristics were detected using SEM, BET, XPS, HRTEM, TEM, FTIR, XRD, bandgap, and VSM. The results showed that with increasing reaction time, radiation intensity, and frequency intensity, the percentage of degradation increased, but with increasing pH and DZN amount, the degradation percentage decreased. The removal efficiency of DZN by the sonophotocatalytic method (100%) was higher than that of the photocatalytic method (89.1%) and the sonocatalytic method (72.9%). The supermagnetic property of the catalyst enables it to be easily separated from the solution and used in different stages of recovery and reuse so that after eight consecutive reuses, only an approximate 9.9% drop was seen in the degradation of DZN. Experiments on scavengers showed that holes and ^•OH radicals play a key role in DZN degradation. The results also showed that the simultaneous use of the photo and sono processes significantly reduces energy consumption, and the reaction rate is much higher than that of separate processes. Finally, results showed that the sonophotocatalytic method is more powerful and leads to more efficient and stable results and the formation of less toxic byproducts, compared to photocatalytic and sonocatalytic methods. As our evaluated system could convert nonbiodegradable into biodegradable wastewater, the usefulness of its employment as a pretreatment process before the biological treatment of industrial or municipal wastewater prior to discharge into the ecological system could be anticipated.

Although the present process showed good activity on a batch scale, during the study of the sonophotocatalytic process, some factors remained undiscovered or were kept constant, which requires further study. For example, the adsorption and oxidation mechanisms can be affected by the speed of mixing and the percentage of materials present in the catalyst, so the effect of changes in the speed of mixing and the percentage of catalyst materials needs to be evaluated. Cost analysis is one of the important factors in the success of a new treatment system, as well as its proposal to improve the solution treatment methods. The feasibility of the new system requires a detailed analysis in terms of cost. In this research, the cost details resulting from the results of the laboratory scale will be misleading. Therefore, in the phase of the pilot scale and full scale, a complete study on cost analysis is suggested. In addition, future studies can consider the treatment of emerging persistent pollutants in real solutions to better understand the sonophotocatalytic process.