Sulfate radicals can be efficiently generated via the activation of PSs, such as peroxydisulfate (PDS, S₂O82-) and permonosulfate (PMS, HSO5-) (Wacławek et al., 2017; Yang W. et al., 2019), both of which are characterized by the presence of an O–O bond (similar to hydrogen peroxide). Generally, PMS, a typical compound salt, mainly consists of potassium hydrogen sulfate, potassium sulfate, and potassium peroxymonosulfate at a ratio of 1:1:2 (Wacławek et al., 2017). In contrast, PDS contains sulfate ions and positive ions.

The asymmetric structure of PMS, generated via the substitution of hydrogen atoms in H₂O₂ by SO₃, is characterized by an O–O bond distance of 1.453 Å (Li et al., 2016). In comparison, PDS possesses a longer bond distance (1.497 Å) and lower bond energy (33.5 kcal mol⁻¹), which is closely related to the substitution of two individual hydrogen atoms of H₂O₂ by SO₃ (Xiao et al., 2018; Song et al., 2019). Thus, theoretically, the chemical characteristics for PDS are more stable than for PMS; PDS also has a higher redox potential than PMS (2.01 vs. 1.77 V) (Wacławek et al., 2017; Xiao et al., 2020). Overall, both PDS and PMS have been widely applied, owing to their excellent water solubility, low cost, ease of storage, and low environmental impact (Xiao et al., 2020). Traditionally, PS refers to PDS, including Na₂S₂O₈, K₂S₂O₈, and (NH₄)₂S₂O₈ (Tsitonaki et al., 2010; Li et al., 2019b). Among them, Na₂S₂O₈ has been frequently applied because of its lower cost, stable chemical characteristics, and ease of solubility.

The activation of PS and PMS is mainly conducted through thermal treatment, UV radiation, use of alkaline, and metal ion reactions. There is a wide variation between sulfate radicals produced by different activation approaches of the PS and PMS, in terms of activation mechanisms, pathways, and redox potential (see Table 1). Thus, summarizing the efficiency of different approaches of PS and PMS is fundamental to listing the critical targets of generation of sulfate radicals (see Figure 1).

Electricity/energy consumption (electricity consumed during the operation of the instruments and equipment) and reagent costs (such as catalytic and oxidation reagents) are the two main costs associated with PS/PMS activation processes. Theoretically, physical activation methods, such as heating, UV, and ultrasonic methods, are relatively economical compared with chemical activation, where the latter incur costs due to reagent consumption. However, physical activation approaches are less applicable in large-scale sewage treatment because of the high energy demand. Taking Fe(II)–PS/PMS systems as an example for organic compound removal from wastewater, only 0.47€ was needed per liter for Fe(II)-activated PS, whereas 4.40€ was needed for UV activation. Although carbon-based materials could be a promising activation method for PS/PMS, their high manufacturing cost restricts their use (even though most biochar is made from wood, sludge, and straw agricultural waste). Hybrid activation always costs more than the other approaches, despite having a higher activation efficiency; thus, further research effort should focus on how to balance activation efficiency and cost.

The degradation of antibiotics, DBPs, and EDCs within the PS/PMS system has been found to be hot spots requiring research attention, and their rate of degradation was recently used to efficiently evaluate their oxidation performance (Wang Q. et al., 2017; Amor et al., 2019; Qin et al., 2020a). Table 2 shows that the degradation efficiency of the different pollutants is closely related to their chemical species/characteristics and the reaction conditions. Generally, the conditions such as temperature, pH, catalyst loading, and concentration of PS/PMS affect the removal rate of substances. For instance, the degradation efficiency of bisphenol A (BPA) was shown to increase gradually (in a linear relationship) from 46.7 to 66.4 to 76.8 to 96.4 and 99.0% with the PS concentrations increased from 7.4 to 15 to 22 to 30 and 35 mM (Xu et al., 2019). However, the degradation rate of ketoprofen (KET) increased from 69.4 to 97.3 to 100% when the pH gradually increased from 3 to 7 to 10, respectively (Feng et al., 2017).

On the other hand, the catalyst activity was shown to significantly influence the degradation of contaminants; this could be proven by the fact that the BPA degradation rate during PMS oxidation with a Sr₂CoFeO₆ catalyst was much higher than that for the reactors without a catalyst additive (Hammouda et al., 2018). The proportion of PS/PMS in the oxidation system was also essential to pollutant degradation because the excess quality of PS can induce a reverse reaction, whereas the insufficient amount induces a low removal rate. The work of Li et al. (2019b) found that the increasing dosage of PS from 1 to 10 mmol/L accelerated the decolorization of reactive brilliant blue (RBB) from 50.42 to 93.75%; however, the excess PS consumed the generated SO4-·, via the following reactions (detailed in Equations 37–38), and led to a significant declining of the overall removal rate of RBB.

Furthermore, the categories of substances also influenced the degradation rate. Under the same experimental conditions (temperature and solution pH), the four typical antibiotics exhibited a decreasing trend—sulfisoxazole (100%) > sulfamethoxazole (98%) ≈ sulfathiazole (97%) > sulfamethizole (81%)—during the oxidation of PS/PMS (Zhou et al., 2019) (see Table 2).

Substitution, decarboxylation, destruction, coupling reaction, etc., accompanied by PS/PMS oxidation, have been shown to lead to distinct intermediate products during contaminant degradation. For instance, the oxidation of bisphenol F resulted in significant production of intermediates of bis(4-hydroxyphenyl)methanol, 4,4′-dihydroxybenzophenone, and 4-hydroxyphenyl 4-hydroxybenzoate as the PS/PMS oxidation progressed (Hammouda et al., 2018). The category of the intermediates is closely related to the degradation pathway and reaction conditions. Wu et al. (2020) found that the molecular weights of the oxidative intermediates of BPA within the Fe₂Co₁-LDH/PS system decreased significantly when the reaction time increased from 10 to 30 to 60 min. The addition of different catalytic components also significantly affects the intermediate products; for instance, the intermediates of BPA in the Fe₃O₄/coal fly ash/PS system were different from those obtained from the CoS@GN-60/PMS system (Zhu et al., 2019). In contrast, the intermediates of BPA in the BC-nZVI/PS system were phenol, p-hydroquinone, 4-isopropenylphenol, and 4-hydroxyacetophenone (Liu et al., 2018a); these are similar to those in the γ-Fe₂O₃@BC/PS system (Rong et al., 2019). Typical intermediates and their structural characteristics of the different containments during the PS/PMS oxidation are summarized in Table 3.

Experiments analyzing toxicity demonstrated that the majority of intermediates were less toxic than their parent pollutant. For example, the toxicity of the detectable intermediates of metolachlor (MET) (Liu C. et al., 2019), landfill leachate (Ghanbari et al., 2020), and carbamazepine (CBZ) (Guo et al., 2020) decreased significantly after PS/PMS oxidation. In contrast, the toxicity of intermediates produced from the degradation of highly concentrated organic wastewater increased significantly because of the formation of some halogenated organics (Wang Q. et al., 2017; Wang et al., 2020); the toxicity of the intermediates of alachlor was comparable to the formation of DBPs after chlorination within the ZVI/PS system (Wang et al., 2020). Similarly, Zhao et al. (2020) found that the formation of chloronitrophenols during thermally activated PS oxidation increased the toxicity of the intermediates during the oxidation of 2-chlorophenol in the presence of nitrite (NO2-) (see Table 3).

Sulfate radicals are critical to the PS/PMS oxidation mechanism; in particular, the kinetic characteristics of the reaction are essential for improving oxidation efficiency. The sulfate radical plays a significant role in pollutant oxidation via the SR-AOP system (see Figure 2 for a summary of the potential mechanisms).

In the PS/PMS system, the degradation efficiency of SO4-· is closely related to the chemical properties of the substrates. For most free radical reactions, SO4-· is the primary reactive oxygen species, and its oxidation ability is 10⁴ times that of SO5-· (Zhou et al., 2018). Specifically, SO4-· can provide electrons to attack the unsaturated bond (–N=C–) of the target pollutant of caffeine (CAF), to form intermediates (Qi et al., 2013). However, the destruction of the benzene ring by SO4-· via hydrogen abstraction is the main pathway for BPA degradation (Zhao et al., 2016).

In addition to the organic substrates, the SO₄⁻· generated can also react within aqueous solutions with sulfite and transition metal ions, as well as itself. Specifically, the most noteworthy sacrificial path of the generated SO₃⁻· pool is the reaction between SO₄⁻· and sulfite; the latter thus has a complicated role in SR-AOPs. Theoretically, increasing the sulfite concentration would improve the substrate oxidation efficiency. However, excessive sulfite will scavenge SO4-· and inhibit substrate oxidation. The concentration and behavior of SO₄⁻· can be analyzed via radical trapping experiments and electron paramagnetic resonance detection (Liu Y. et al., 2019; Wu et al., 2020). Further oxidation of H₂O or OH⁻ by SO₄⁻ leads to the generation of OH· (Equations 39–40), which would be another key reason for pollutant oxidation. The SO4-· and OH· produced can oxidize organic pollutants to different intermediates or inorganics through various degradation pathways (Hammouda et al., 2018). For example, Qi et al. (2013) found that the generated SO4-· and OH· radicals, with the Co²⁺/PMS system, can not only efficiently degrade caffeine but also mineralize the corresponding intermediates (see Figure 2).

The high concentration of sulfates, generated from the PS/PMS activation processes and side reactions, posed a potential threat to aquatic life (limited concentration of sulfate is 500 mg/L). However, the production and discharging of the sulfates during the PS/PMS activation are seldom focused. The recent studies revealed that one of the efficient approaches to overcome this limitation is the development of integrating the SR-AOP technique with cheaper oxidants, e.g., hydrogen peroxide or ozone (Duan et al., 2020). For example, Yang et al. (2015) proposed that the generated sulfates could be efficiently oxidated by the reaction of ozone with PMS.

SR-AOPs are complicated physicochemical processes; studying the thermodynamic and kinetic characteristics of the reaction of different pollutants is essential for clarifying the mechanisms. Generally, the pseudo-first-order (Equation 39) and pseudo-second-order kinetics (Equation 40) have been widely used for the simulation of the degradation reaction of organic compounds; these can be described by linear forms in Equations (41) and (42), respectively.

Where C₀ is the initial concentration of the pollutant, C_t represents the pollutant concentration at a certain reaction time (t), and k₁ and k₂ are the pseudo-first-order and pseudo-second-order rate constants, respectively.

The respective species and chemical structures of organic pollutants significantly affect the kinetic characteristics of degradation. Some of the latest literature published revealing the kinetic mechanisms of typical organic pollutants during SR-AOPs is presented in Table 4. Overall, the degradation processes of (1) typical EDCs (such as bisphenol A, bisphenol F, and tetrabromobisphenol A) (Ji et al., 2016b; Hammouda et al., 2018; Xu et al., 2019) and (2) PPCPs (such as ketoprofen, ibuprofen, carbamazepine, etc.) (Feng et al., 2017; Wang et al., 2019; Guo et al., 2020) fit well with the pseudo-first-order kinetics. In contrast, the recent work of Lutze et al. (2015) found that the pseudo-second-order model yielded a much better fit for simulating the oxidation process of the chlorotriazine pesticides (such as atrazine, desethylterbuthylazine, and terbuthylazine).

It is interesting to note that the degradation of bisphenol S in a heat/PS system followed a pseudo-zero-order model (Wang Q. et al., 2017), whereas neither of the abovementioned zero-order, first-order, and pseudo-second-order models could be applied for the simulation of the degradation of diclofenac within the bismuth ferrite (BFO)/PMS system. Han et al. (2020) stated that the diclofenac degradation curve could be well-fitted to a double exponential expression, as shown in Equation (43).

Where C_t and C₀ are the diclofenac concentration (mM) at times t and 0, respectively; a₁ and a₂ represent fraction coefficients and k₁ and k₂ are the apparent rate constants (min⁻¹) for the first and second types of diclofenac degradation, respectively.

The reaction rates of different organic pollutants differ widely. Taking the reaction rate constant obtained from pseudo-first-order kinetics, for example, the abovementioned PPCPs, EDCs, and HCOM decrease in a trend of PPCPs (10⁻²-10⁻¹) ≈ EDCs (10⁻²-10⁻¹) > HCOM (10⁻⁴-10⁻³). Specifically, the obtained k₁ value of ketoprofen was 0.26 min⁻¹ in the heat/PS system at 70°C, whereas it declined to 3.29 × 10⁻² min⁻¹ when the temperature decreased to 60°C. The order of magnitude of the reaction rate constants was about 10⁻³-10⁻⁴ for pseudo-second-order kinetics (Khan et al., 2017). In addition, the changes in the experimental conditions could also affect the reaction rate constant; for example, the k₁ value of bisphenol A during the reaction significantly increased from 0.07 to 0.26 min⁻¹, when the catalyst loading increased from 0.2 to 0.4 g/L (Rong et al., 2019).

The species of radicals generated during SR-AOPs would also significantly affect the decomposition rate constant of the target pollutants. Zhao et al. (2016) reported that the degradation rate constant of BPA could be attributed to both the reactions with SO4-· and OH·, respectively, which can be described by Equations (44).

Where k′ represents the second-order rate constant for the reaction of BPA with each individual oxidizing species. Similarly, Wang et al. (2016) also found that the rate constants for the reaction of OH· were higher than those for SO4-· during the degradation process of 2,4-di-tert-butylphenol. Moreover, Nasseri et al. (2017) found that the second-order rate constants of DEA [(6.42 ± 0.12) × 10⁸ M⁻¹ s⁻¹] and DIA [(1.70 ± 0.30) × 10⁹ M⁻¹ s⁻¹] under SO4-· oxidation were lower than those of OH· [k_DEA = (1.14 ± 0.09) × 10⁹ M⁻¹ s⁻¹, k_DIA = (2.22 ± 0.44) × 10⁹ M⁻¹ s⁻¹] (see Table 4).

As the dominant radical, SO4-· oxidizes the pollutants primarily by the following mechanisms: (1) hydrogen abstraction, (2) electron transfer, and (3) addition or substitution reaction. Electron transfer is the predominant mechanism of the reaction between SO4-· and aromatic organics. Taking the SR-AOP degradation of, for example, ketoprofen (Feng et al., 2017), bisphenol F (Hammouda et al., 2018), and bisphenol A (Xu et al., 2019), the attack or destruction of the benzene ring by SO4-· might be the main reason for their degradation. By comparison, hydrogen abstraction could easily occur during SO4-· reaction with saturated organic compounds (e.g., alkanes, alcohols, organic acids, ethers, and esters) (Zhao et al., 2016). For alkene and alkyne organics, the single electrons of the sulfate radical would attack the unsaturated bonds, and addition or substitution would enhance the degradation.

Generally, the concurrence of the abovementioned degradation pathways plays a significant role in the degradation of organic chemicals. For instance, the electrophilic attack of SO4-·, hydroxylation, and coupling reactions were the main reasons for BPA oxidation within the ZVI/PS system (Zhao et al., 2016). In contrast, a deprotonate radical attacking an unsaturated bond in oxidizing and substitution reactions belonged to the CAF during the application of the Co-MCM41/PMS system (Qi et al., 2013). The non-free radical reaction was another reason for the partial pollutant degradation. For instance, the ¹O₂ generated from the self-decomposition of PMS (via Equation 45) would attack pollutants/intermediates via an electron transfer mechanism. According to the recent work of Liu Y. et al. (2019), ¹O₂ could easily be generated either from the reactions between ·O2- and OH·/H⁺ or from direct formation from SO5-· (see Figure 3).