RESEARCH PAPER

 

Degradation of methyl orange by the Fenton-like reaction of pyrite-activated hydrogen peroxide forming the Fe(lll)/Fe(ll) cycle

 

 

Wenlong Bi; Ruojin Du; Hui Liu; Peng Fu; Zhenguo Li

College of Resources and Environment, Shanxi Agricultural University, Shanxi, 030801, PR China

Correspondence

 

 

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ABSTRACT

In this study, the typical azo dye methyl orange (MO) was degraded by pyrite (FeS₂) activated by hydrogen peroxide (H₂O₂). When [MO] = 0.1 mM, [FeSJ = 2.0 g/L and [H₂O₂] = 22 mM, 96.4% MO was removed in 120 min and the TOC removal rate was higher than 50%. HO* was the primary radical responsible for MO degradation, in addition, the acid condition promoted the degradation of MO in the FeS₂/H₂O₂ system. MO in tap water and river water was not effectively degraded, whereas acidification could weaken the inhibitory effect on the FeS₂/ H₂O₂ system to enable the degradation of MO in tap and river water. The OD₆₀₀ indicated that the solution was environmentally friendly after the reaction, and three degradation pathways of MO were discussed. In summary, Fe(ll) could be dissolved from FeS₂, which activated H₂O₂ to generate Fe(lll) and HO*. FeS₂ could reduce Fe(lll) into Fe(ll), thus realizing the Fe(lll)/(ll) cycle and efficiently activating H₂O₂ to degrade MO.

Keywords: pyrite, hydrogen peroxide, methyl orange, advanced oxidation processes

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INTRODUCTION

With the rapid development of printing and dyeing textile industries, the large amount of printing and dyeing wastewater produced has become one of the main water pollution sources in the world (Gu et al., 2021; Pinedo-Hernández et al., 2022). Printing and dyeing wastewater has a complicated composition, high concentration of pollutants, poor biochemical properties, and is difficult to treat (Ben Hafaiedh et al., 2020; Domingues et al., 2019; Shukla and Remya, 2021). Among the contaminants, methyl orange (MO) contains -N=N- and is a typical azo dye (Aghdasinia et al., 2016; Liu et al., 2022). Wastewater containing MO is discharged into the environment from industrial processes such as textiles and printing, and without effective treatment can have a serious impact on the environmental safety of surrounding water and soil, endangering human health (Giannakis et al., 2021). The accumulation of MO in water reduces the penetration of sunlight, inhibits the growth of aquatic organisms, and leads to an imbalance in the ecosystem (Song et al., 2022). Therefore, the study of efficient treatment methods for methyl orange wastewater has great significance.

Advanced oxidation processes (AOPs) are used to oxidize macromolecular refractory organics into environmentally friendly small molecules by generating free radicals with strong oxidizing properties through electricity, ultraviolet light, and ultrasound (Liu et al., 2022; Peng et al., 2019; Shao et al., 2020). AOPs, such as the Fenton reaction which involves reacting Fe²⁺ with H₂0₂ to produce hydroxy] radical (HO'), have a strong oxidation capacity, a fast reaction rate, high mineralization potential, and simple operation (Almeida et al., 2021; Gorozabel-Mendoza et al., 2021; Zhao et al., 2020). However the Fenton system has a high catalyst dosage and low pH after reaction (Jin et al., 2021). In addition, the high iron content of FeSO₄ added to the Fenton system greatly increases the iron sludge treatment intensity and the disposal cost (Guo et al., 2020; Martins et al., 2017). Therefore, finding a low-cost and environmentally friendly catalyst to replace Fe(II) is critical for advanced oxidation processes.

To solve the iron-rich sludge problem in the Fenton system, a cheap and easily available iron-containing mineral needs to be found to replace ferrous ions. A promising candidate is pyrite (FeS₂), which is the most abundant sulfide mineral on earth. FeS₂ has been shown to be an efficient catalyst for the degradation of alachlor by Fenton's reagent (Kayan et al., 2020), and a nano-pyrite Fenton system was demonstrated to have improved the production of HO', resulting in the degradation of p-nitrophenol being three times faster than with the traditional Fenton system (Liu et al., 2020). Moreover, FeS₂ is a low-cost, easily available, and environmentally friendly catalyst (Wang et al., 2021). The most noteworthy attribute of FeS₂ is that the resource utilization of FeS₂ can reduce the risk of FeS₂ producing acidic mine wastewater (AMD) (Xia et al., 2023).

Therefore, in this study, FeS₂ was used as a catalyst to explore the degradation and mechanism of MO in FeS₂/H₂O₂ system. Furthermore, considering that printing and dyeing enterprises use a lot of tap water and river water, the degradation of MO in distilled water, tap water, and river water in FeS₂/ H₂O₂ system was investigated.

 

MATERIALS AND METHODS

Chemicals

Methyl orange (MO, C₁₄H₁₄N₃SO₃Na, >98.8%) was purchased from Tianjin Tianli Chemical Reagent Co., Ltd. Hydrogen peroxide (H₂0₂, >30%) and concentrated sulfuric acid (H₂SO₄, 95.0%~98.0%) were purchased from Chengdu Kolon Chemical Reagent Co., Ltd. Ferrous sulfate (FeS0₄-7H₂O, >99.0%), sodium hydroxide (NaOH, 96.0%), methanol (CH₃OH, >99.5%), and tert-butyl alcohol (TBA, (CH₃)₃COH, >99.5%) were purchased from Tianjin Beichen Founder Reagent Factory. All the above reagents were analytically pure. All aqueous solutions were prepared with distilled water.

Pyrite pretreatment

Natural pyrite (FeS₂) was purchased from Hebei China. Pyrite was ground and passed through a 0.15 mm mesh sieve, placed in a beaker and soaked in 0.5 M H₂SO₄ solution for 1 h to remove impurities and then filtered. The pyrite was washed repeatedly with distilled water until the pH of the filtrate was 7, and then filtered. The filtered sample was dried in an oven at about 40°C for future use.

Experimental method

MO removal by pyrite alone or hydrogen peroxide alone

FeS₂ (0.2-2.0 g/L) or H₂0₂ (30-50 mM) was added to MO (0.1 mM) and the solution placed on a magnetic stirrer; 3 mL samples were taken at 5,10,20, 30, 60, 90, and 120 min and then to each sample 0.5 mL of MeOH was added, which can quench free radicals to ensure sample stability. The concentration of MO was calculated by colorimetric measurement at 470 nm (iso-absorption point wavelength, which is the wavelength at the intersection of the acidic and alkaline absorption curves) with a 721 visible spectrophotometer.

Optimization of pyrite and hydrogen peroxide

The addition of FeS₂ and H₂O₂ was optimised using the response surface method in Design-Expert software. The design of experiment (DOE) is given in Table 1 which was designed using response surface methodology (RSM) (Ebru et al., 2023). The initial concentration of MO was 0.1 mmol/L, and the software-designed doses of FeS₂ and H₂O₂ were added sequentially before the start of the reaction. 3 mL of the sample was removed from the filtrate at 15, 30, 45 and 60 min by passing it through a 0.45 µm membrane and adding 0.5 mL of methanol (free radical scavenger), and the concentrations were measured using a 721 visible spectrophotometer. The pH and temperature of the solution was measured at the end of the reaction. The data were visualised and analysed in Design-Expert software to obtain the optimum dose of FeS₂ and H₂O₂.

 

 

Effect of different factors

In pH-effect experiments, the pH of the MO (0.1 mM) solution was adjusted to 2, 4, 6, 8 and 10 with H₂S0₄ and NaOH, respectively and then FeS₂ and H₂O₂ were added after the above optimization. The reaction was started and timed at 5, 10, 20, 30, 60, 90 and 120 min. 3 mL of the sample was removed from the filtrate after passing through a 0.45 µm filter membrane and 0.5 mL of methanol was added. The concentrations were measured using a 721 visible spectrophotometer, and the pH and temperature of the solution were measured at the end of the reaction. In the free-radical-contribution experiment, 2 mL TBA was added to MO solution, after adding optimized FeS₂ and H₂0₂ into the MO solution; the remaining experimental steps were as above. All water used in the above procedures was distilled water. The influence experiment for different water matrices was conducted using distilled water, tap water, and river water (the Uma River in Taigu) to prepare MO solutions for degradation experiments, and the remaining experimental steps were performed as above.

Biotoxicity experiments

Biotoxicity experiments used microbial communities obtained by isolation and activation from campus soil (the soil was taken to cultivate the microorganisms used in the experiment in the area where the vegetation on the campus was lush and the microorganisms were active). The composition and content of the medium is given in Table 2. The 15 mL reaction solution was added to 10 mL of the medium and 0.5 mL mixed microbial community solution, and placed in an oscillator (rotating speed 180 r/min, temperature 25°C) for oscillating culture. Samples were taken at 6, 9, 12, 24 and 48 h, and the OD₆₀₀ value reflecting bacterial density was determined by spectrophotometer. According to the difference between the reaction solution and the blank treatment, the microbial toxicity of the treatment solution was determined.

 

 

Analysis methods

Determination of total organic carbon concentration was conducted using a total organic carbon analyser (TOC; Muti N/C-3100 model). A scanning electron microscope (SEM; JSM-7001F) and X-ray diffractometer (XRD; MiniFlex II) were used to characterize FeS₂. FeS₂ surface elements were analysed using Al-Ka monochromatic X-ray photoelectron spectroscopy (XPS; AXIS ULTRA DLD). The intermediate products of MO degradation were analysed by a liquid chromatography-mass spectrometer (LC-MS; Agilent 1290 11-6470).

 

RESULTS AND DISCUSSION

Removal of MO by FeS₂ alone and H₂O₂ alone

The removal effect for MO by different concentrations of FeS₂ was examined as shown in Fig. 1a. When the FeS₂ concentration was in the range of 0.2-2.0 g/L, the adsorption capacity of FeS₂ for MO was weak and fluctuated up and down in the range of 0-3.5%. With an increase in FeS₂ concentration, adsorption of MO did not improve effectively. Figure 1b shows the removal effect for MO by different concentrations of H₂O₂. The removal rates of MO at 30, 40 and 50 mmol/L H₂O₂ after 150 min were 2.8%, 5.0% and 4.9%, respectively. As the concentration of H₂O₂ increased, removal of MO did not increase significantly, indicating that the oxidation of MO by H₂O₂ was weak.

The degradation of MO in FeS₂/H₂O₂ with different concentrations of H₂O₂

As can be seen from Fig. 2, when 0.1 mmol/L MO and [2.0 g/L FeS₂ were added, the residual of MO at 30, 40 and 50 mmol/L H₂O₂ after 150 min of reaction was 97.4%, 93.5% and 96.1%, respectively. The degradation rate of MO was significantly increased after the simultaneous addition of FeS₂ and H₂O₂. This may be due to the continuous dissolution of Fe(II) from FeS₂ and the continuous activation of H₂O₂ by Fe(II) to produce IIO' (Eq. 1) to degrade MO (Oral and Kantar, 2019). Moreover, the sulfide in FeS₂ could reduce Fe(III) to Fe(II) (Eq. 2) (Liu et al., 2015), further promoting the Fenton-like reaction. With the increase in H₂O₂ concentration, the difference between the removal rates for methyl orange was only 3.9%. Excluding the interference of error lines and combined with the desire for green and efficient' a relatively low concentration of H₂O₂ (30 mmol/L) was finally chosen for the subsequent optimization experiments.

 

 

 

Determination of the optimal concentration of FeS₂ and H₂O₂

Response surface method optimization

By analysing the results of MO removal by FeS, alone, H,0, alone, and FeS,/H₂O₂ system (Figs 1 and 2), the concentration ranges of FeS₂ and H₂O₂ were determined, and then 3D surface plots of MO removal rate after 150 min were fitted by Design-Expert software for the interaction between FeS₂ and H₂O₂ concentrations (Abd Manan et al. 2019). As can be seen from Fig. 3, when the MO concentration was 0.1 mM, H₂O₂ concentration was 22 mM, and FeS₂ concentration was 2.0 g/L, the MO removal rate was expected to reach 100% after 150 min.

 

 

The MO removal rate increased with the increase in FeS₂ concentration. This was probably due to the degradation of MO with the increased FeS₂ that could dissolve more Fe(II)-activated H₂O₂ generating HO^- production.

When the MO concentration was 0.1 mM, FeS, concentration was 2.0 g/L, and with the increase in H₂O₂ concentration, the MO removal rate showed a trend of first increasing and then decreasing. The MO removal rate increased with the increase in H₂O₂ concentration in the range of 0-22 mM, which may be due to Fe(II) dissolved from FeS₂, then reacting with H,0, to generate HO'. However, MO removal decreased once the H₂O₂ concentration was above 22 mM. On the one hand, H₂O₂ is a capture agent for HO', and excess H₂O₂ would not only react with Fc(II), but also with HO^- to produce HO,' with low activity (Eq. 3) (Chang et al., 2020; Sun et al., 2019; Xiao et al., 2020), making the reaction rate decrease. On the other hand, excess HO' would react not only with the target substance but also with Fc(II), which made the degradation rate of MO decrease.

Verification of optimal conditions

To verify the above response surface methodology optimization results, the MO removal by different systems was plotted in Fig. 4a. MO removal by FeS₂ or H₂O₂ alone was minimal. After 150 min of reaction, the MO removal by FeS₂ alone was only 3.5%, and FeS₂ had no obvious oxidative adsorption effect on MO.

The MO removal rate by H₂O₂ alone was only 4%, so the oxidation effect of H,0, on MO was weak.

When the MO concentration was 0.1 mM, H,0, concentration was 22 mM, and FeS, concentration was 2.0 g/L, the MO removal rate reached 96.3% at 120 min. The MO removal rate showed a trend of plateau, then a high rate, and finally a plateau again, which is consistent with a traditional Fenton-hke trend (Wang et al., 2013; Yan et al., 2017). The rapid degradation of MO at the stage of 30-90 min may be due to Fc(II) being dissolved from FeS₂, which then reacted with H,0, to produce HO". FeS, also reduced Fe(III) to Fe(II), which continues to react with H₂O₂, resulting in rapid degradation of MO. The results for the UV scanning spectrum at different times also indicated rapid degradation of MO within 30-90 min (Fig. 4b). During the 90-150 min stage, the degradation of MO was slow. On the one hand, H₂O₂ was gradually consumed in FcS₂/H₂O₂ system (Kokilavani et al., 2021), which made only a small amount of H₂O₂ react with Fe(II). On the other hand, intermediates of MO would compete with MO for IIO', slowing the MO degradation rate. To sum up, compared with FeS₂ alone or H₂O₂ alone, the FeS₂/H₂O₂ system could significantly remove MO. In addition, the degradation of MO tends to level off after 120 min. Because the concentration of MO at 150 min remained almost unchanged compared to that at 120 min, the final sampling time for subsequent experiments was 120 min.

Table 3 shows MO removal rates by other activated H₂O₂ methods. In Fc(II)/H₂O₂ system, although the MO is degraded in a short time, the amount of Fe(II) is large and the cost is high.

In UV/H₂O₂ system, an additional light source is needed and a large amount of H₂O₂ is required. Compared with other iron-hased/H₂O₂ systems, FeS₂/H₂O₂ system could efficiently degrade 96.3% of MO with low doses of FeS, and H₂O₂. In addition, FeS₂ is a mineral that exists in nature, is low cost and easy to obtain, and has good application value.

Contribution of different factors to FeS₂/H₂O₂ system

Concentration of total Fe and Fe(ll)

Fe(II) plays a catalytic role in FeS,/H₂O₂ system, which could activate H₂O₂ to generate HO^- and effectively degrade MO, and all iron ions in this system originate from FeS,. Therefore, we measured the total Fe and Fe(II) concentrations in FeS₂/H₂O₂ system as shown in Fig. 5a. The total Fe concentration increased with time, but all of the concentrations recorded were lower than 3.5 mg/L, The Fe(II) concentration fluctuated from 0.33-0.64 mg/L. After 30 min of reaction, the total Fe and Fe(II) concentrations were 0.43 mg/L and 0.33 mg/L respectively, indicating that the slow degradation of MO in FeS₂/H₂O₂ system before 30 min was due to the slow dissolution of Fe ions. During 30-120 min, the total Fe content increased significantly, reaching 3.12 mg/L at 120 min, while Fe(II) did not increase, reaching 0.64 mg/L at 120 min. This maybe due to the rapid reaction of dissolved Fe(II) with H₂O₂ to form Fe(III), FeS₂ could reduce Fe(III) to Fe(II), and the reaction rate of Eq. 1 is faster than that of Eq. 2 (Diao and Chu 2021). Therefore, there was a large amount of Fe(III), and Fe(II)/Fe(III) in dynamic states, in the system, confirming that the degradation of MO was rapid after 30 min in FeS₂/H₂O₂ system.

Validation of the interface effect of pyrite

H₂O₂ and Fe(II) were added to the MO solution to form Fe(II)/H₂O₂ homogeneous system to investigate the interface effect of FeS₂. It can be seen from Fig. 5b that MO degraded slowly in Fe(II)/H₂O₂ system, and the degradation rate of MO was 77.4% after 120 min. H₂O₂ rapidly produced a large amount of HO' under the action of a high concentration catalyst in the early stage (Dulova et al., 2017). However, due to the limited reaction rate of HO' with organic matter, the free HO' not participating in the reaction aggregates, and free HO' reacted with itself to form H₂O₂ resulting in a part of the initially generated HO' being wasted (Liu et al., 2021). Furthermore, there were a large number of intermediate products of MO in the system, which might compete with MO for HO'. Therefore, the degradation trend of MO tended to level off. The pH after the reaction of the Fe(II)/H₂O₂ and FeS₂/H₂O₂ systems was 4.00 and 3.25, respectively. FeS₂ has the advantages of low cost and easy availability, so it has good application prospects.

Free radical contribution

To investigate the reactive oxygen species that played a dominant role in the removal of MO from FeS,/H₂O₂ system, a certain amount of free radical scavenger was added to the reaction system. The reaction rate constants of tert-butanol (TBA) and HO' are (3.8-7.6) x 10^s M/s, so TBA is commonly used as a HO' scavenger (Chen et al. 2021, Xu and Sheng 2021). In addition, to effectively quench HO' within the system, we added TBA (0.4 M) at 4 000 times the dose of MO (0.1 mM). The effect of TBA on the degradation of MO in FeS₂/H₂O₂ system is shown in Fig. 5c. The addition of TBA significantly inhibited the degradation of MO, and the MO removal rate was only 40.2% in 120 min. Therefore, HO' was the dominant reaction species in the system, and its contribution to the degradation of MO was about 58.3%. The interfacial effect of FeS, also contributed as the added FeS₂ particle interface reacted with H₂O₂.

Influence of pH on the degradation of MO in FeS₂/H₂O₂ system

The degradation of MO in FeS₂/H₂O₂ system under different pH levels was shown in Fig. 6. In FeS₂/H₂O₂ system, MO could be effectively removed under neutral and acidic conditions, and the removal effect was better under acidic conditions (Jin et al., 2021). However, it was difficult to effectively remove MO at pH - 10.

When the initial pH was 2 and 4, the MO degradation showed a fast and then slow trend. The degradation rates in 0-60 min were fast, and the MO removal rate could reach 94.8% and 87.5% in 60 min. The degradation tends to reach equilibrium within 60-120 min, with an increase of 2.4% and 8.7% in MO removal at 120 min compared with 60 min. Under acidic conditions, the dissolution of Fe(II) in FeS₂ could be promoted (Li et al. 2019). Therefore, in the early stage of the reaction, the Fe(II) would

immediately activate H₂O₂, which made MO degrade rapidly. The slow degradation of MO after 60 min may be due to the decrease of H₂O₂, and only a small amount of Fe(II) reacted with a small amount of H₂O₂, which made MO degrade slowly. When the initial pH was 6 and 8, MO removals were 94.0% and 93.0%, respectively, after 120 min. The degradation trends were all flat, then fast and finally flat again, similar to the degradation trends at unadjusted pH (Dos Santos et al., 2021). The rapid degradation of MO in 30-90 min was not only due to the massive dissolution of Fe(II) to produce HO', but also because FeS, would produce H⁺ during the reduction of Fe(III), resulting in more dissolution of Fe(II) after the reduction of pH. Under the alkaline condition, the MO removal rate was poor at the initial pH of 10, and was only 2.2% after 120 min. It may be that Fe(II) dissolved slowly in FeS, under the alkaline condition (Dong et al., 2019).

When the initial pH was 10, the pH after the reaction was 9.71 However, with all of the other starting pH values, the pH after the reaction was less than 4 (Table 4). During the reduction of Fe(III), FeS₂ produces H⁺, which lowers the pH of the solution. Under a traditional Fenton system, the best degradation efficiency was achieved at pH 2.79. At an initial pH of 8.0, there was almost no degradation of MO (Youssef et al., 2016). Compared with the traditional Fenton system, the FeS₂/H₂O₂ system has a wider range of pH applications.

Influence of different water matrices on MO degradation in FeS₂/H₂O₂ system

In practice, most printing and dyeing textile industries would directly use tap water or river water for treatment and, after a series of processes, the discharged wastewater has a complex composition. Therefore, we investigated MO degradation in tap water and river water by FeS₂/H₂O₂ system to verify the possibility of FeS, treatment toward disposal of actual printing and dyeing wastewater.

As shown in Fig. 7a, MO was not effectively degraded in river water and tap water, and its degradation effect was much lower than that in distilled water. By determining the physical and chemical properties of river water and tap water (Table A1, Appendix) it was shown that they were weakly alkaline in pH and all had high TOC content. Consequently, it is speculated that the alkaline water would react with HO' to form low levels of active free radicals (Eqs 4-5) (Jiang et al., 2021; Ma et al., 2018), and the organic matter in river water and tap water would compete with MO for HO^-. Hence there was no obvious removal effect for MO.

Under the acidic condition, the dissolution of Fe (II) will be accelerated and this will promote MO degradation. Moreover, the acidification could remove CO₃^2-, HCO₃^-, Br^-, NO₃^-, and Cl^- in water. The distribution of carbonation binding states in Fig. A1 (Appendix) shows that at pH < 4 there is only free CO₂ in the water. Therefore, the effect of different water matrices at pH = 2 was further investigated (Fig. 7b). After 60 min of reaction, the MO removal rates in distilled water, tap water, and river water were 97.2%, 96.9%, and 97.6%, respectively, achieving the expected treatment effect. Moreover, we prepared MO wastewater with similar ions to the water taken from the Uma River. It was found

that the simulated wastewater was difficult to degrade in FeS₂/H₂O₂ system without adjusted pH, whereas the simulated wastewater had good degradation at a pH of 2 (Fig. 8), which further confirmed that acidification could weaken the inhibitory effect on the FeS₂/H₂O₂ system for degrading MO in tap and river water.

Characterization of pyrite before and after reaction

The elemental composition of FeS, and the chemical states of Fe and S elements before and after the reaction were analysed by XPS. As can be seen from Fig. 9a, FeS, before and after the reaction was mainly composed of S, C, O, Fe, and Na. It could also be detected by EDS spectra (Fig. A2, Appendix). The C and O may be surface oxides, and Fe and S decreased in intensity after use, which may be as a result of being dissolved from the mineral into water. A corresponding change in the relative intensity of FeS, before and after the reaction could also be observed by XRD (Fig. 9b).

The high-resolution Fe-2p spectra of FeS, is shown in Fig. 9c. The peaks of 707.3, 709.6, 710.8, 711.5, and 714.2 eV of Fe 2p_3/² orbit were FeS₂, FeO, Fe₂O₃, FeOOH, and satellite peaks, respectively. Except for the satellite peak, the peak area ratios before reaction were 40.2%, 10.7%, 23.6%, and 25.5%, respectively. The FeS, peak area decreased by 8.9%, while the FeOOH peak area increased by 8.7% after the reaction.

As can be seen from Fig. 9d, the peaks at 162.6 and 163.9 eV correspond to the S 2p_3/2 and S 2p_1;2 orbitais of S,^2- (Ye et al, 2021) The peak area of S₂² decreases from 77.7% to 54.6% before and after the reaction. The peaks at 168.5 and 169.7 eV are the S 2p_3/2 and S 2p_1/2 orbitais of SO₄² , and the peak area of SO₄² increased by 23.1% before and after the reaction. This indicates that S,² in FeS, was oxidized to S0₄^2- during the oxidation process, which was consistent with the reaction given in Eq. 2 (Gao et al., 2021) Since the conversion of elemental S was much larger than that of elemental Fe, it was speculated that sulfide in FeS, could promote the conversion of Fe(III) to Fe(II) (Yu et al., 2020).

Determination of biotoxicity and total organic carbon of MO degraded by FeS₂/H₂O₂ system

The growth of microorganisms was observed by determining the OD₆₀₀ value to assess the biotoxicity of the FeS₂/H₂O₂ system's degradation of MO. As can be seen from Fig. 10a, during 0-12 h, the number of microorganisms in the FeS₂/H₂O₂ system was more than that in distilled water and MO solution. However, the biotoxicity of FeS₂/H₂O₂ system at 48 h was not significantly different from that of MO, further suggesting that the FeS₂, H₂O₂ and intermediate degradation products of MO were environmentally friendly. Thus, there is an excellent application prospect.

The total organic carbon (TOC) concentration in the FeS,/H₂O₂ system was measured to observe the degradation of MO (Suave et al., 2018). The TOC removal rate increased with increasing reaction time (Fig. 10b), indicating that the total carbon content in MO wastewater gradually decreased with time (Hamad et al., 2016). In FeS₂/H₂O₂ system, the TOC removal rate reached 58.8% after 2 h of reaction. In summary, the FeS,/H₂O₂ system was capable of effectively degrading MO.

Possible degradation pathways and mechanism of MO in FeS₂/H₂O₂ system

To determine the degradation path of MO, LC-MS was used to analyse the degradation intermediates of MO. Figure A3 (Appendix) shows the liquid-phase mass spectra of MO at 120 min; 8 intermediates of MO were inferred, and their molecular structures are listed in Table A2 (Appendix). The degradation pathways are shown in Fig. 11a. In FeS₂/H₂O₂ system, Na of MO generally existed in the form of ion, and MO degradation paths were roughly divided into two: one was demethylation and deaminization for PI, P5, and P6 products; P7 was the -N=N-breakage product (Zhang et al., 2020), and P3 and P8 were deaminization products. Pathway 2 was that MO first broke the N=N- to P2, further -NH-NH- broke to P3, P4, and eventually became other small molecules. Figure 11b shows the degradation mechanism for MO in FeS,/H₂O₂ system. Fe(II) dissolved from FeS₂, and Fe(II) activated H₂O₂ to generate HO", while sulfide in FeS₂, could reduce Fe(III) to Fe(II) and continuously activate H₂O₂ to generate HO" for MO degradation.

 

CONCLUSION

In this study, H₂O₂ activated by FeS₂ was effective in degrading MO and HO' was the main radical. In addition, the removal rate of TOC was higher than 50%. In addition, the reaction solution was environmentally friendly and three possible pathways for the degradation of MO were analysed. Compared with the traditional Fenton system, the FeS₂/H₂O₂ system has a wider range of pH applications. Furthermore, MO in tap water and river water could not be effectively degraded in FeS₂/H₂O₂ system. However, acidity could promote MO degradation; hence MO in tap water and river water at pH = 2 was rapidly degraded within 60 min. The degradation mechanism was that Fe(II) dissolved from FeS₂, Fe(II) activated H₂O₂ to generate HO', which could degrade MO. Meanwhile, FeS₂ could reduce Fe(III) to Fe(II). Therefore, the Fe(III)/Fe(II) cycle formed could continuously activate H₂O₂ to produce HO'to degrade MO.

 

AUTHOR CONTRIBUTIONS

The manuscript was written by Hui Liu. Ruojin Du contributed to the preparation and review of this manuscript. Experiments were performed by Peng Fu, and data analyses were carried out by Zhenguo Li. The project was supervised by Wenlong Bi. All authors have read and agreed to the published version of the manuscript.

 

ACKNOWLEDGEMENTS

This work was supported by the Youth Foundation of Shanxi Agricultural University Science and Technology Innovation Enhancement Project (CXGC2023026). Thanks to the staff of the College of Resources and Environment, Shanxi Agricultural University, and we are grateful to the anonymous reviewers and the editors for their valuable suggestions and comments on revising and improving this paper.

 

CONFLICT OF INTEREST

The authors declare no conflict of interest.

 

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Correspondence:
Wenlong Bi
Email: jiayangbihu@126.com

Received: 14 May 2023
Accepted: 22 March 2024

 

 

APPENDIX

 

 

 

Table A2 - Click to enlarge

 

 

 

 

 

 

Figure A3 - Click to enlarge