Introduction

Sulfentrazone, a protoporphyrinogen oxidase (PPO)-inhibiting herbicide, has been widely used for controlling broad-leaved weeds and some grasses (e.g., pigweed, waterhemp, and crabgrass) in soybean, sugarcane, and tobacco [1]. However, sulfentrazone is highly stable, and the degradation of sulfentrazone in the environment is very slow. Its half-life extension in soil could reach 541 d [2]. The solubility of sulfentrazone in water is 780.0 mg/L [3]. It can easily enter natural water bodies through rainfall and surface runoff leading to water contamination [4, 5]. It was documented that sulfentrazone has a high risk to aquatic plants. Kong et al. found that sulfentrazone is highly toxic to Scenedesmus obliquus with 72 h-ErC50 of 0.415 mg a.i./L and 72 h-EyC50 of 0.227 mg a.i./L, respectively [6]. Also, sulfentrazone can induce neurologic disorders in vertebrates and affect bacterial growth [7, 8]. Therefore, limiting sulfentrazone residue is crucial for minimizing its risks to the environment.

Traditionally, three major methods are used to reduce sulfentrazone contamination in water, including adsorption [9], electrochemical advanced oxidation [10] and microbial degradation [11]. Diogo et al. reported the Carbon nanostructures supported on Co/serpentinite is effective in adsorbing sulfentrazone in water [12]. However, most the porous adsorbents are unlikely to exhibit desirable adsorption capacity and desorption efficiency [9, 13]. Lima et al. adopted an electro-Fenton method for degrading sulfentrazone in water and achieved a mineralization of 60.0% for 200.0 mg/L sulfentrazone in 4 h [14]. However, the electrochemical degradation method may incur a high cost and toxic secondary metabolites [10]. Melo et al. isolated six bacterial strains for degrading sulfentrazone in water, but the degradation rate of sulfentrazone only ranged from 4.0% to 15.0% within 10 d [15]. Microbial degradation is mostly confined to the laboratory stage due to the poor survival ability of bacteria and fungi, as well as fluctuating residue degradation rates [16].

The activation of persulfate such as peroxymonosulfate (PMS) and persulfate (PS) using carbon-based materials for removing organic pollutants has been regarded as an efficient and friendly advanced oxidation processes (AOPs) in recent years [17, 18]. Compared to AOPs via semiconductor photocatalysis [19], sulfate radicals (SO4•–) are produced from activating persulfate in addition to hydroxide radicals (e.g., O2•−, 1O2 and •OH) generated by catalytic reactions [20]. SO4•– possesses a longer half-life and stronger redox ability compared to hydroxide radicals, demonstrating more efficient oxidative activity in the degradation of organic pollutants [Sulfentrazone degradation

The sulfentrazone degradation procedure was as follows: first, 15.0 mg of GM-HC and 5.0 mg of PMS were added to 50.0 mL of 10.0 mg/L sulfentrazone, followed by shaking at 120 rpm at 25 °C. At set intervals, 1.0 mL of solution was collected, and the reaction was quenched immediately by adding 1.0 mL of MeOH. To determine the effect of various factors on sulfentrazone degradation by PMS activated with GM-HC, the initial concentration of sulfentrazone (2.50–40.0 mg/L), PMS dose (0.50–6.0 g/L), GM-HC amount (0.50–4.0 g/L), initial pH (3.35–8.51) adjusted using H2SO4 or NaOH solutions, reaction temperature (15–45 °C), and concentration of inorganic anions (Cl−, HCO3−, H2PO4−, NO3−, SO42−) were evaluated. Three recycling runs of GM-HC were carried out to evaluate its reusability under the same reaction conditions. The concentration of remaining sulfentrazone was determined by high-performance liquid chromatography (HPLC), and the degradation intermediates and products of sulfentrazone were detected by UPLC-QTOF/MS after filtration with a 0.22 µm filter.

Analysis of active species

To identify the active species generated during the GM-HC activation of PMS, we conducted quenching experiments using various scavengers. MeOH at 10.0, 50.0, or 100.0 mM was used to quench SO4•– and •OH; TBA at 10.0, 50.0, or 100.0 mM was used for •OH; p-BQ at 5.0, 10.0, or 20.0 mM was used for O2•−, and FFA at 5.0, 10.0, or 20.0 mM was used for 1O2 [30]. A certain amount of scavenger was added to a 100.0 mL glass beaker containing 10.0 mg/L sulfentrazone, 3.0 g/L GM-HC and 1.0 g/L PMS by shaking at 120 rpm at 25 °C. At 0, 1, 8, 12, 16, and 24 h, 1.0 mL of solution was added to 1.0 mL of MeOH. The sample underwent filtration using a 0.22 µm filter, and the amount of sulfentrazone was measured using HPLC.

HPLC and UPLC-QTOF-MS analysis

The amount of sulfentrazone remaining following degradation was analyzed using an HPLC system (Agilent LC1260, USA) with a Kromasil® C18 column (250.0 mm × 4.60 mm, 5.0 µm). Sulfentrazone was detected by an ultraviolet detector at 240.0 nm with a mobile phase (MeOH:water containing 0.10% acetic acid), 70:30, v/v) at a flow rate of 0.80 mL/min. The degradation products of sulfentrazone were analyzed by UPLC-QTOF-MS (Agilent UPLC1290-QTOF6530, USA) by ESI negative ion injection voltage with an ACCHROM XAqua C18 column (150 mm × 4.6 mm, 5 μm) under the following conditions: mobile phase (0.20% acetic acid solution in water and acetonitrile) with a flow rate of 0.30 mL/min and a column temperature of 35 °C. The scanning range (m/z) was 50–500 amu.

DFT calculations

DFT calculations were used to predict the reactive sites for nucleophilic, electrophilic, and radical attacks based on Fukui functions [31]. The geometric conformation of the sulfentrazone molecule was established using GaussView 5.0.8. Then, geometric optimization and frequency analysis were performed using Gaussian 09 at the B3LYP/6-31G* level. Frontier electron density (FED) describes the electronic properties and reactivity of molecules and predicts the contribution rates of the highest occupied molecular orbital (HOMO) and the lowest unoccupied molecular orbital (LUMO) to identifying reaction sites in molecules [32]. The frontier electron density of the HOMO (\({FED}_{HOMO}^{2}\)) and the frontier electron density of the LUMO (\({FED}_{LUMO}^{2}\)) were calculated using the method described by Lee et al. [33].

Detoxification bioassay

To investigate the toxicity caused by sulfentrazone and detoxification of sulfentrazone in rice by GM-HC/PMS, rice was selected as a target for toxicity assessment because it is very sensitive to sulfentrazone. Bioassays were conducted in Petri dishes for rice seed germination and in plastic cups for rice growth. All dishes and cups were put into a growth chamber at 25/28 °C night/d under a 10 h-dark and 14 h-light cycle with a 12,000 lx light intensity and 80.0% relative humidity. For the seed germination bioassays, a total of 30 seeds were evenly placed in a 9-cm diameter Petri dish. Two pieces of filter paper were placed in a dish and moistened with 5.0 mL of one of three aqueous solutions (distilled water, solution with concentrations of 0.20, 1.0, and 5.0 mg/L of sulfentrazone, or solution of sulfentrazone treated with 3.0 g/L of GM-HC coupled 1.0 g/L of PMS). Germination was monitored for 5 d and the percentage of seeds that germinated seeds was recorded (Eq. (1)). For the rice growth bioassays, a total of 10 germinated rice seeds were exposed to 5 mL of one of the three aqueous solutions described above. After 5 d of growth, the length of the rice shoots was measured.

$$\mathrm{Germination}\;\mathrm{rate}\;(\%)\;=\;(\mathrm{germinated}\;\mathrm{seeds}/\;\mathrm{total}\;\mathrm{number}\;\mathrm{of}\;\mathrm{seeds})\;\times\;100$$
(1)

Toxicity analysis

The toxicity of sulfentrazone and its intermediates was predicted by the Toxicity Evaluation Software Tool (T.E.S.T., 19.0.0.0) program.

Data analysis

The data were presented as means ± standard deviations of three replicates. The significance of differences between the treatments was assessed using an independent-samples t-test, *P < 0.05.

Results and discussion

Effect of PMS concentration on sulfentrazone removal

The effect of the PMS concentration on sulfentrazone removal was studied over a range of 0.50–6.0 g/L. The 24 h sulfentrazone removal efficiency increased from 76.30% to 86.30% as the PMS concentration increased from 0.50 g/L to 1.0 g/L (Fig. 1A). However, with further increase of the concentration of PMS to 6.0 g/L, the removal efficiency was reduced to 72.0%. Enhancing PMS concentration generates more sulfate radicals, leading to an increase in the sulfentrazone removal efficiency [34, 35]. However, SO4•− and •OH may also be scavenged through undesirable reactions when the PMS concentration is high, resulting in a decrease in sulfentrazone removal efficiency that we observed at concentrations > 1.0 g/L (Eqs. (2) and (3)) [36, 37].

Fig. 1
figure 1

Effect of main experimental factors on sulfentrazone removal efficiency. A PMS concentration; (B) GM-HC dosage; (C) Initial concentration of sulfentrazone; (D) Initial solution pH; (E) Temperature; (F) Common anions

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$$\bullet \text{OH }+\text{ HS}\text{O}_{5}^{-}\to \text{SO}_{5}\bullet^{-} + {\text{H}}_{2}\text{O}$$
(2)
$$\text{SO}_4\bullet^-+\text{ HS}{{\text{O}}_{5}}^{-}\rightarrow\text{ S}\text{O}_5\bullet^-+\text{ HS}{{\text{O}}_{4}}^{-}$$
(3)

Effect of the dose of GM-HC on sulfentrazone removal

Sulfentrazone removal was measured at diverse GM-HC dosages. The GM-HC dose significantly affected sulfentrazone removal (Fig. 1B). The removal efficiency of sulfentrazone was enhanced when the GM-HC dose was increased from 0.50 g/L to 4.0 g/L. Higher amounts of GM-HC resulted in higher removal efficiency because of an increase in the number of active sites available for PMS decomposition to produce more active species (e.g., SO4•−, O2•−, and •OH) [17, 38, 39]. When the GM-HC dose was increased from 3.0 g/L to 4.0 g/L, the removal efficiency of sulfentrazone reached 100.0% during a 24-h reaction with the GM-HC/PMS system. It was reported that an excessive amount of catalyst can lead to aggregation, hindering the reaction between pollutants and free radicals [40]. However, Ding et al. found that increasing the concentration of hydrochar-montmorillonite from 0.60 g/L to 3.0 g/L could catalyze PMS to produce more reactive species, thereby enhancing the removal efficiency of dicamba [17], which is consistent with our empirical data in this study.

Effect of initial sulfentrazone concentration on removal

The initial concentration of sulfentrazone significantly affected its removal efficiency (Fig. 1C). The removal efficiency of sulfentrazone decreased as its concentration increased. After 24 h, the removal efficiencies of sulfentrazone at concentrations of 2.5, 5.0, 10.0, 20.0, and 40.0 mg/L were 100.0%, 100.0%, 93.90%, 73.0%, and 52.60%, respectively. Because the PMS dose remained fixed as the sulfentrazone concentration increased, this fixed PMS dose was not sufficient to completely remove sulfentrazone from the system at higher concentrations [41]. Ma et al. observed that the removal efficiency of imidacloprid was negatively correlated with its initial concentration in the G-HWTRs/PMS system [42], and Ding et al. reported similar results for quinclorac removal in the nZVI/ATP-PMS system [43].

Effect of initial pH on sulfentrazone removal

The removal efficiencies of sulfentrazone were 95.61%, 95.51%, 94.20%, 94.10%, and 93.10% after 24 h of reaction at initial pH values of 3.35 (unadjusted pH), 5.25, 6.84, 7.35, and 8.51, respectively (Fig. 1D). This result indicates that the removal efficiency of sulfentrazone is minimally influenced by the pH levels in acidic, neutral, and alkaline conditions when employing the GM-HC/PMS system, enabling effective operation across a broad pH range [44]. These findings are in line with those reported by Qi et al. [45].

Effect of temperature on sulfentrazone removal

Sulfentrazone removal efficiency was positively correlated with temperature in the GM-HC/PMS system (Fig. 1E). 8 h sulfentrazone removal efficiency increased from 64.25% to 100.0% when the reaction temperature was elevated from 15 °C to 45 °C. Temperature is a crucial factor in the activation of PMS. Li et al. found that the efficiency of ofloxacin degradation by heat-activated persulfate was dramatically enhanced with increased temperature [46]. Ding et al. also demonstrated that the degradation of quinclorac was significantly accelerated with increasing temperature in the presence of nZVI/ATP3 activating PMS [43]. Thus, temperature could contribute to the diffusion rate of PMS and accelerate generation of SO4•– and •OH, enhancing removal efficiency [47].

Effects of anions on sulfentrazone removal

The addition of anions significantly influenced the sulfentrazone removal efficiency (Fig. 1F). The removal efficiencies of sulfentrazone were 56.85%, 94.25%, 34.05%, 97.21% and 97.29% after 24 h of reaction involving Cl−, HCO3−, H2PO4−, NO3−, and SO42−, respectively, and were 93.32% without the addition of anions. This result indicates that the presence of Cl− and H2PO4− decreased the removal efficiency of sulfentrazone, perhaps because the generated SO4•− and •OH could react with Cl−, as demonstrated in Eqs. (4)-(8), and H2PO4−, as demonstrated in Eq. (8), to yield fewer reactive radicals [34, 48, 49]. The unexpected phenomenon of the removal efficiency of sulfentrazone lower by adding H2PO4− reaction than by adding Cl− might be due to the reason that excess Cl− reacted with PMS and Cl• to produce reactive halogens such as Cl2•− and HOCl, which also could oxidize sulfentrazone according to Eqs. (4)-(8). In contrast, HCO3−, NO3− and SO42− had little effect on the removal efficiency of sulfentrazone by the GM-HC/PMS system [50].

$$\text{HS}{{\text{O}}_{5}}^{-}+{\text{Cl}}^{-}\to \text{S}{{\text{O}}_{4}}^{2-}+\text{HOCl}$$
(4)
$$\text{HS}{{\text{O}}_{5}}^{-}+2{\text{Cl}}^{-}+{\text{H}}^{+}\to \text{S}{{\text{O}}_{4}}^{2-}+{\text{Cl}}_{2}+{\text{H}}_{2}\text{O}$$
(5)
$$\text{SO}_{4}\bullet^{-}+\text{Cl}^{-}\to \text{Cl}\bullet+\text{S}{{\text{O}}_{4}}^{2-}$$
(6)
$$\text{Cl}\bullet+\text{ C}{\text{l}}^{-}\to \text{Cl}_{2}\bullet^{-}$$
(7)
$$\bullet \text{OH}+\text{H}_{2} \text{PO}_{4}^{-}\to \text{OH}^{-}+\text{H}_{2} {\text {PO}}_{4} \bullet$$
(8)

Reusability of GM-HC

To assess the reusability of GM-HC activating PMS for removing sulfentrazone, sulfentrazone degradation was carried out three successive cycles at the same reaction condition. Figure 2 showed the sulfentrazone removals for three consecutive runs. As observed, the degradation efficiency of sulfentrazone gradually decreased with increasing reuse time. The degradation efficiency of sulfentrazone decreased from 97.16% to 24.44% after the third cycle. A possible reason for this decrease is the deposition of degradation products on the surface of the GM-HC, leading to a decrease in the number of available reaction sites for activating PMS [51].

Fig. 2
figure 2

Effect of the reuse times of GM-HC on sulfentrazone removal efficiency

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Determination of active species

The active species generated during sulfentrazone removal in the GM-HC/PMS system were identified using several scavengers [52]. The results showed that the four scavengers MeOH, TBA, p-BQ, and FFA had adverse effects on the sulfentrazone removal efficiency in the GM-HC/PMS system, and this inhibition of sulfentrazone removal became more pronounced as the scavenger concentration increased (Fig. 3). Sulfentrazone removal efficiency ranged from 32.83% to 57.96% when MeOH was added to the reaction system, indicating that SO4•− and •OH play crucial roles in the degradation of sulfentrazone (Fig. 3A). Sulfentrazone removal efficiency ranged from 70.61% to 87.68% when TBA was used to quench •OH, suggesting that SO4•− contributes more to sulfentrazone removal than •OH (Fig. 3B). Both p-BQ and FFA had significant inhibitory effects on sulfentrazone removal efficiency, reducing the removal efficiency to below 25.0%, suggesting that large amounts of O2•− and 1O2 were involved in these reactions (Fig. 3C, D). These combined results suggest that SO4•−, O2•−, and 1O2 are the dominant active species in sulfentrazone degradation. These species may be attributed to the oxygen-containing functional groups, including doublet aliphatic C-OH and C = O and phenolic C-O/Si-O, on the surface of GM-HC transferring electrons to PMS, causing the cleavage of the peroxide O-O bond in PMS to form SO4•−, which further generates O2•− and 1O2. The reaction of glucose-montmorillonite hydrochar composite activating PMS facilitates sulfentrazone degradation [17].

Fig. 3
figure 3

Effects of scavengers on sulfentrazone removal. A Methanol (MeOH); (B) Tertbutyl alcohol (TBA); (C) 1,4-benzoquinone (p-BQ); (D) Furfuryl alcohol (FFA)

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Degradation products and pathway of sulfentrazone

The degradation intermediates and products were detected by UPLC-QTOF/MS. The mass spectrum of sulfentrazone and its degradation intermediates were displayed in Supplementary Materials Fig. S1A-F. Sulfentrazone (C11H10Cl2F2N4O3S, m/z 384.98) and its five degradation intermediates were identified as 1-(2,4-dichloro-5-(methylsulfonamido)phenyl)-4-(difluoromethyl)-5-oxo-4,5-dihydro-1-H-1,2,4-triaz-ole-3-carboxylic acid (S1,C11H8Cl2F2N4O5S, m/z 414.82), N-(2,4-dichloro-5-(4-(difluoromethyl)-3-(hydroxyl-methyl)-5-oxo-1,2,4-triazolidin-1-yl)phenyl)methanesulfonamide (S2,C11H12Cl2F2N4O4S, m/z 402.80), N-(4-chloro-3-(4-(difluoromethyl)-3-methyl-5-oxo-4,5-dihydro-1H-1,2,4-triazol-1-yl)phenyl)methanesulfonamide (S3,C10H8Cl2F2N4O3S, m/z 370.81), N-(3-(3,4-dimethyl-5-oxo-4,5-dih-ydro-1H-1,2,4-triazol-1-yl)phenyl) methanesulfonamide (S4, C11H11ClF2N4O3S, m/z 351.02), and 4-(difluorom-ethyl)-5-methyl-2,4-dihydro-3H-1,2,4-triazol-3-one (S5, C11H14N4O3S, m/z 281.04) (Supplementary Materials Table S1).

Based on the chemical structure of sulfentrazone and its metabolites, two potential degradation pathways of sulfentrazone in the GM-HC/PMS system have been suggested (Fig. 4). In the first proposed pathway, the chlorine atoms on the benzene ring of the sulfentrazone molecule were removed by attacking •OH, leading to the generation of intermediate S4 followed by S5. The dechlorination of hydrogen atoms substituting for chlorine atoms has been documented in the degradation pathway of quinclorac [53]. After the dechlorination of sulfentrazone, S5 is oxidized by the active species into small inorganic molecules, CO2, and H2O. In the second proposed pathway, a hydrogen atom on the methyl group (-CH3) in the triazol ring is initially removed under the energy of pπ-conjugation to form a methylene radical (-CH2•), which binds with •OH to produce hydroxymethyl (-CH2OH), leading to the generation of the metabolite S2. Then, the -CH2OH group at the triazol ring of sulfentrazone undergoes oxidation to a carboxyl group (–COOH) upon attack by SO4•−, 1O2, and O2•−, leading to the formation of product S1 [54, 55]. Subsequently, a decarboxylation reaction occurs to produce S3 [17, 56]. Considering these two pathways, it is highly plausible that sulfentrazone undergoes stepwise degradation involving substitution, oxidation, decarboxylation, and dechlorination, ultimately mineralizing into CO2, H2O, and small inorganic molecules.

Fig. 4
figure 4

Proposed degradation pathways of sulfentrazone in GM-HC/PMS system

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Quantum chemistry analysis of sulfentrazone degradation

To further demonstrate the degradation products obtained from UPLC-QTOF/MS and the degradation pathway of sulfentrazone, quantum chemical analyses of sulfentrazone were performed based on the optimal geometric conformation of the sulfentrazone molecule obtained at the B3LYP/6-31G* level. The spatial configuration of the sulfentrazone molecule is presented in Fig. 5A. As an effective approach for predicting the degradation products and pathways of organic compounds, DFT calculations based on the HOMO, LUMO, and electrostatic potential (ESP) surface, \({\text{FED}}_{\text{HOMO}}^{2}\) and \({\text{FED}}_{\text{LUMO}}^{2}\) have been broadly employed to explain the reactivity sites on organic molecules [24, 54]. The HOMO represented the most electron-rich regions at C 4, C 6, Cl 7, C1 8, and N 17, and the LUMO represented electron-deficient sites at C 4, C 6, C1 8, and N 17 (Fig. 5B, C). Moreover, the FED values of the HOMO and LUMO indicate that C 4 and C 6 possess the highest FEDHOMO2 and FEDLUMO2, with values of 0.2506 and 0.3650, respectively (Table 1).

Fig. 5
figure 5

DFT calculations on sulfentrazone molecule. A Optimized chemical structure of sulfentrazone; (B) The highest occupied molecular orbital (HOMO); (C) The lowest unoccupied molecular orbital (LUMO); (D) Electrostatic potential (ESP)-mapped molecular surface of sulfentrazone

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Table 1 Frontier electron density of highest occupied molecular orbital (HOMO) and lowest unoccupied molecular orbital (LUMO) on major atoms of sulfentrazone at the B3LYP/6-31G* level
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Figure 5 showed that atoms C 4 and C 6 represent vulnerable sites attacked by radicals during sulfentrazone degradation, leading to cleavage of C 6-Cl 7 and C 4-Cl 8 and the generation of degradation products S4 and S5. This theoretical research on frontier electron densities was reported by Wang et al. to elucidate the degradation of indomethacin [57]. Figure 5D shows the ESP, highlighting the susceptibility of C 4, Cl 7, Cl 8, N 17, C 18, and O 20 to attack by electrophilic free radicals. These sites are easier for active species to attack and exhibit high degradation activity [30, 58]. The quantum analyses aligned with the degradation products identified through UPLC-QTOF/MS and the suggested degradation pathway. These findings provide valuable insights into the reactivity of sulfentrazone, aiding in a comprehensive understanding of its degradation behavior in the GM-HC/PMS system.

Detoxification of sulfentrazone by GM-HC/PMS

Previous studies have shown that sulfentrazone can cause serious phytotoxicity to sensitive crops [59, 60]. To evaluate the detoxification effect of the GM-HC/PMS system against sulfentrazone-induced phytotoxicity to rice, rice germination and seedling growth were investigated. The results showed that sulfentrazone significantly suppressed rice seed germination (Fig. 6A) and seedling growth (Fig. 7A) due to herbicidal phytotoxicity. The degree to which phytotoxicity was alleviated by the GM-HC/PMS treatment varied similarly between the rice germination and seedling growth groups. Under exposure to sulfentrazone at 0.20, 1.0, and 5.0 mg/L, the percentage of germinated seeds increased by 14.67%, 34.67%, and 89.33%, respectively, in the GM-HC/PMS treatment group compared to that in the untreated control group (Fig. 6B), and the shoot length increased by 19.42%, 67.86%, and 73.05%, respectively (Fig. 7B). The rice seed germination rate and seedling height were not significantly different in GM-HC/PMS-treated plants relative to those kept in distilled water. Moreover, it was also found that both 3.0 g/L GM-HC and 1.0 g/L PMS were benificial to plant growth. This phenomenon of good to crop growth was documented in previous research. Ding et al. observed that 3.0 g/L of hydrochar-montmorillonite (HC-Mt) and 0.61 g/L of PMS promoted the growth of mung bean [17]. Wen et al. utilized 12.0 g/L of biochar (WCGBC800) activating 60.0 g/L of PMS to alleviate the inhibitory effects of quinclorac on rice. They discovered that WCGBC800 and PMS promoting seed germination and seedling growth [61]. There was few reports on the overuse of catalysts and PMS affecting plant growth.

Fig. 6
figure 6

Rate of rice seed germination in sulfentrazone solution with treatment by the GM-HC/PMS system after 5 days growth. Note: Sul: sulfentrazone. Asterisks indicate significant differences: *, P < 0.05, 0.01 by Student’s t-test

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Fig. 7
figure 7

Effect of sulfentrazone with/without GM-HC/PMS treatment on shoot length of rice seedlings. Note: Sul: sulfentrazone. Asterisks indicate significant differences: *, P < 0.05, 0.01 by Student’s t-test

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To comprehensively assess the ecological risks of the GM-HC/PMS system, the ecological toxicity of sulfentrazone and its degradation intermediates was predicted based on the Toxicity Estimation Software Tool (T.E.S.T.) [53]. The results are shown in Table 2. The results revealed that the toxicity of the degradation intermediates to Daphnia magna and the fathead minnow was lower than that of sulfentrazone, indicating that the GM-HC/PMS system can detoxify sulfentrazone to aquatic organisms, which is consistent with the alleviation of sulfentrazone-induced phytotoxicity to rice. Regarding the remediation efficiency of existing methods for sulfentrazone degradation and phytotoxicity alleviation in rice, the adsorption of biochar and multi-walled carbon nanotubes was able to alleviate the phytotoxicity of sulfentrazone to rice [62], but residual sulfentrazone remains in the environment after desorption. Similarly, Co2+ and Fe2+/PMS System were also used to alleviate the phytotoxicity of sulfentrazone to rice, but metal ions leaching causes secondary environmental pollution [63]. These combined results illustrate that the GM-HC/PMS system is a viable alternative for sulfentrazone detoxification.

Table 2 The toxicity of sulfentrazone and its degradation intermediates to Daphnia magna and Fathead minnow calculated by T.E.S.T software program
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Conclusions

GM-HC catalyzes PMS to accelerate the degradation of sulfentrazone in water. When GM-HC was 3.0 g/L and PMS was 1.0 g/L, a removal efficiency of 10.0 mg/L for sulfentrazone reached 93.90% within 24 h. The main reason is that the SO4•−, O2•−, and 1O2 generated by GM-HC/PMS contributed to rapid degradation. Five degradation intermediates of sulfentrazone were detected, and two pathways were proposed, which were confirmed by DFT calculations. In addition, a phytotoxicity bioassay verified that GM-HC/PMS could alleviate the effects of sulfentrazone toxicity in rice seeds and seedlings. Our combined results indicate that GM-HC, which activates PMS, is a feasible alternative for addressing pesticide residues and their phytotoxicity to crops.