A Review of Manganese(III) (Oxyhydr)Oxides Use in Advanced Oxidation Processes

The key role of trivalent manganese (Mn(III)) species in promoting sulfate radical-based advanced oxidation processes (SR-AOPs) has recently attracted increasing attention. This review provides a comprehensive summary of Mn(III) (oxyhydr)oxide-based catalysts used to activate peroxymonosulfate (PMS) and peroxydisulfate (PDS) in water. The crystal structures of different Mn(III) (oxyhydr)oxides (such as α-Mn2O3, γ-MnOOH, and Mn3O4) are first introduced. Then the impact of the catalyst structure and composition on the activation mechanisms are discussed, as well as the effects of solution pH and inorganic ions. In the Mn(III) (oxyhydr)oxide activated SR-AOPs systems, the activation mechanisms of PMS and PDS are different. For example, both radical (such as sulfate and hydroxyl radical) and non-radical (singlet oxygen) were generated by Mn(III) (oxyhydr)oxide activated PMS. In comparison, the activation of PDS by α-Mn2O3 and γ-MnOOH preferred to form the singlet oxygen and catalyst surface activated complex to remove the organic pollutants. Finally, research gaps are discussed to suggest future directions in context of applying radical-based advanced oxidation in wastewater treatment processes.


Introduction
Over the past few decades, with the rapid development of industrialization and the increase of anthropogenic activities, huge amounts of organic and inorganic contaminants were discharged into the surface and ground waters, causing water pollution problems and threatening human health [1][2][3]. However, conventional water treatment technologies, such as filtration [4,5], precipitation [6,7], coagulation-flocculation [8][9][10], and biological treatment [11,12] exhibited a minimal effect on the removal of recalcitrant pollutants. Therefore, there is an increasing demand for efficient, economical, and environmental-friendly water treatment technologies. Advanced oxidation processes (AOPs) have attracted particular attention due to their high efficiency for removal of recalcitrant contaminant. AOPs are able to remove and mineralize most unbiodegradable pollutants into harmless compounds, such as CO 2 , H 2 O, and inorganic ions [13]. Based on various reaction conditions, AOPs can be classified into different categories, including Fenton reaction [14], Fenton-like reaction [15,16], photochemical oxidation [17,18], ultrasonic oxidation [19,20], electrochemical oxidation [21,22], ozone oxidation [23,24], and sulfate radical-based AOPs (SR-AOPs) [25][26][27]. Among them, the application of SR-AOPs for the removal of stubborn pollutants has received increasing attention due to their advantages. For instance, sulfate radical (SO •− 4 ) has a longer lifetime compared with the hydroxyl radical (HO • ), a wide range of pH adaptation, and a high reduction potential (2.5-3.1 V vs. NHE) [28].
Generally, the peroxydisulfate (PDS, S 2 O 2− 8 ) and peroxymonosulfate anions (PMS, HSO − 5 ) are employed as the radical precursors for producing sulfate radicals through The authors also demonstrated the effects of structured MnO 2 on peroxymonosulfate (PMS) activation, and the low reactivity of δ-MnO 2 was attributed to its less crystallinity [43]. Crystalline manganese oxides are generally built on the same basic unit [MnO 6 ] octahedral with the edges or corners sharing [41]. The commonly reported Mn(III) (oxyhydr)oxides include manganese(III) oxide (α-Mn 2 O 3 ), groutite (α-MnOOH), feitknechtite (β-MnOOH), manganite (γ-MnOOH), and hausmannite (Mn 3 O 4 ). The structures covered in the name of Mn(III) (oxyhydr)oxides are summarized in Table 1. Among them, α-Mn 2 O 3 , γ-MnOOH, and Mn 3 O 4 have attracted increasing attention from the scientific community because of their promising technological applications, such as in catalysis, water treatment, and ion exchange. The crystalline structure of α-Mn 2 O 3 was recognized as the body-centered cubic bixbyite phase, as shown in Figure 1a. γ-MnOOH possesses a typical (1 × 1) tunnel structure constructed by [MnO 6 ] octahedral sharing the corners (Figure 1b). The structure of γ-MnOOH is analogous to that of pyrolusite, except that one-half of the oxygen atoms are replaced by hydroxyl anions compared with pyrolusite. For the crystalline Mn 3 O 4 , it exhibits a normal spinel structure with the formula Mn 2+ (Mn 3+ ) 2 O 4 where the Mn 2+ and Mn 3+ ions occupy the tetrahedral and octahedral sites, respectively ( Figure 1c). currence of more accessible active sites in layered δ-MnO2 than other tunnel structured MnO2 [49]. The authors also demonstrated the effects of structured MnO2 on peroxymonosulfate (PMS) activation, and the low reactivity of δ-MnO2 was attributed to its less crystallinity [43]. Crystalline manganese oxides are generally built on the same basic unit [MnO6] octahedral with the edges or corners sharing [41]. The commonly reported Mn(III) (oxyhydr)oxides include manganese(III) oxide (α-Mn2O3), groutite (α-MnOOH), feitknechtite (ꞵ-MnOOH), manganite (γ-MnOOH), and hausmannite (Mn3O4). The structures covered in the name of Mn(III) (oxyhydr)oxides are summarized in Table 1. Among them, α-Mn2O3, γ-MnOOH, and Mn3O4 have attracted increasing attention from the scientific community because of their promising technological applications, such as in catalysis, water treatment, and ion exchange. The crystalline structure of α-Mn2O3 was recognized as the body-centered cubic bixbyite phase, as shown in Figure 1a. γ-MnOOH possesses a typical (1 × 1) tunnel structure constructed by [MnO6] octahedral sharing the corners (Figure 1b). The structure of γ-MnOOH is analogous to that of pyrolusite, except that one-half of the oxygen atoms are replaced by hydroxyl anions compared with pyrolusite. For the crystalline Mn3O4, it exhibits a normal spinel structure with the formula Mn 2+ (Mn 3+ )2O4 where the Mn 2+ and Mn 3+ ions occupy the tetrahedral and octahedral sites, respectively ( Figure  1c).  The influence of structures in the reactivity of common Mn(III) (oxyhydr summarized in Table 2. For instance, Saputra et al. investigated the effect of mo on the oxidation of phenol by Mn2O3 activated PMS. The results showed that cub has the highest reactivity on PMS activation in comparison with octahedral-and octahedral-Mn2O3, and it was due to the high surface area and distinct surface rangement of cubic-Mn2O3 [55]. Similarly, Cheng et al. successfully prepared Mn2O3 in cubic-, truncated octahedral-, and octahedral-structure, and investigat fect of crystal facets on the combustion of soot [56]. The results show that the s bustion efficiency followed the order of α-Mn2O3-cubic > α-Mn2O3-truncated o > α-Mn2O3-octahedral. The enhanced reactivity of α-Mn2O3-cubic was explain fact that the exposed (001) surface facets of α-Mn2O3-cubic have higher amoun coordinated surface oxygen sites, which are capable of facilitating the oxygen a and improving the surface redox properties.
In addition to α-Mn2O3, it was also reported that the oxidative and catalyt mances of Mn3O4 and γ-MnOOH were affected by their structures. For exampl reported that the hexagonal nanoplate Mn3O4 exhibited superior catalytic perfor diesel soot combustion compared to the octahedral and nanoparticle Mn3O4, and ing was explained by the improved amount of surface Mn 4+ species and surfac oxygen species due to the increased fraction of exposed (112) facets in hexago plate Mn3O4 [57]. The effect of morphology was also discovered by Liu et a demonstrated that the nanoflake Mn3O4 (exposure of (001) facet) has the highe reduction reactivity in comparison to nanoparticle Mn3O4 and nanorod Mn3O4 ( of (101) facet) [58]. In addition, He et al. investigated the activation of PMS by γ with different shapes, and the results showed that the catalytic activity of γ-Mn lowed the order of nanowires > multi-branches > nanorods [59]. Different The influence of structures in the reactivity of common Mn(III) (oxyhydr)oxides is summarized in Table 2. For instance, Saputra et al. investigated the effect of morphology on the oxidation of phenol by Mn 2 O 3 activated PMS. The results showed that cubic-Mn 2 O 3 has the highest reactivity on PMS activation in comparison with octahedral-and truncated octahedral-Mn 2 O 3 , and it was due to the high surface area and distinct surface atoms arrangement of cubic-Mn 2 O 3 [55]. Similarly, Cheng et al. successfully prepared three α-Mn 2 O 3 in cubic-, truncated octahedral-, and octahedral-structure, and investigated the effect of crystal facets on the combustion of soot [56]. The results show that the soot combustion efficiency followed the order of α-Mn 2 O 3 -cubic > α-Mn 2 O 3 -truncated octahedral > α-Mn 2 O 3 -octahedral. The enhanced reactivity of α-Mn 2 O 3 -cubic was explained by the fact that the exposed (001) surface facets of α-Mn 2 O 3 -cubic have higher amounts of lowcoordinated surface oxygen sites, which are capable of facilitating the oxygen activation and improving the surface redox properties.
In addition to α-Mn 2 O 3 , it was also reported that the oxidative and catalytic performances of Mn 3 O 4 and γ-MnOOH were affected by their structures. For example, Ji et al. reported that the hexagonal nanoplate Mn 3 O 4 exhibited superior catalytic performance on diesel soot combustion compared to the octahedral and nanoparticle Mn 3 O 4 , and the finding was explained by the improved amount of surface Mn 4+ species and surface reactive oxygen species due to the increased fraction of exposed (112) facets in hexagonal nanoplate Mn 3 O 4 [57]. The effect of morphology was also discovered by Liu et al., which demonstrated that the nanoflake Mn 3 O 4 (exposure of (001) facet) has the highest oxygen reduction reactivity in comparison to nanoparticle Mn 3 O 4 and nanorod Mn 3 O 4 (exposure of (101) facet) [58]. In addition, He et al. investigated the activation of PMS by γ-MnOOH with different shapes, and the results showed that the catalytic activity of γ-MnOOH followed the order of nanowires > multi-branches > nanorods [59]. Different physicochemical parameters, such as specific surface area, Lewis sites, zeta-potential, and redox potential were measured to study the reason for the different catalytic performances of γ-MnOOH with distinct morphologies. It was found that the charge density on the surface played a crucial role in the interfacial reactivity between PMS and γ-MnOOH. In summary, the reactivity of Mn(III) (oxyhydr)oxides on radical precursor activation and pollutant oxidation can be deeply affected by their structures. The desirable morphologies and facets (such as cubic structure with (001) facet exposure) can apparently improve the reactivity of Mn(III) (oxyhydr)oxides.

Activation of PMS by Mn(III) (Oxyhydr)Oxides
The Mn(III) (oxyhydr)oxides/PMS system has been applied for the removal of a number of contaminants, such as phenol, bisphenol A, 2,4-dichlorophenol, ciprofloxacin, and organic dyes [62][63][64][65][66][67]. Different studies involving PMS activation by Mn(III) (oxyhydr)oxides are gathered in Table 3. According to the literature, the efficient degradation of organic pollutants is generally attributed to the generation of active species, such as SO •− 4 , HO • , 1 O 2 . The activation mechanisms of PMS by Mn(III) (oxyhydr)oxides are proposed, as shown in Figure 2. The simultaneous formation of Mn(II) and Mn(IV) and the conversion of Mn ions with different oxidation states explained well the good performance of Mn(III) (oxyhydr)oxides on PMS activation (Equations (1)-(4)) [44]. Except for the abovementioned processes, the direct generation of HO • by Mn(III) activation of PMS was also reported by some researchers (Equation (5)) [62,64,66,[68][69][70]. In comparison with SO •− 4 radical, the SO •− 5 radical has been regarded as a low oxidative activity for organic pollutants removal due to its low reduction potential (E 0 = 1.10 V vs. NHE) [71]. Nevertheless, the transformation from SO •− 5 to SO •− 4 in Mn(III) (oxyhydr)oxides/PMS system still makes some contribution to the degradation of organic pollutants (Equation (6)) [72]. In addition, the conversion from SO •− 4 to HO • in water should not be neglected (Equation (7)), especially, when the solution is in the alkaline environment (Equation (8)) [73].
ules 2021, 26, x FOR PEER REVIEW makes some contribution to the degradation of organic pollutan addition, the conversion from SO 4 • to HO • in water should no (7)), especially, when the solution is in the alkaline environment In addition to the active radicals, the generation of non-radical species (such as 1 O 2 ) in the Mn(III) (oxyhydr)oxides-activated PMS system was also reported.  [76].
Currently, the Mn-based oxide composites have attracted increasing attention due to their various advantages, such as more oxygen vacancies, higher surface oxygen mobility, and enforced synergistic effects. For instance, Chen et al. prepared the Fe 2 O 3 /Mn 2 O 3 composite and studied its activity on PMS activation for tartrazine (TTZ) degradation. The results showed that 97.3% removal of TTZ was achieved in 30 min in the Fe 2 O 3 /Mn 2 O 3 /PMS system. The efficient degradation of TTZ originated from the generation of active species (e.g., SO •− 4 , HO • ) and the synergistic effect between iron and manganese ions [77]. The γ-MnOOH-coated nylon membrane was synthesized and applied in the activation of PMS towards the removal of 2,4-dichlorophenol (2,4-DCP). The deep removal of 2,4-DCP was explained by the synergetic "trap-and-zap" process, which improved the stability and catalytic reactivity of γ-MnOOH [63]. In conclusion, the activation of PMS by Mn(III) (oxyhydr)oxides, including pure Mn(III) oxides and Mn(III) containing composites, is favorable. The degradation of various pollutants in the Mn(III) (oxyhydr)oxides/PMS system can be achieved through the generation of active radicals and non-radical species.

Activation of PDS by Mn(III) (Oxyhydr)Oxides
Single or combined Mn(III) (oxyhydr)oxides have been employed to activate PDS to remove different organic pollutants, such as phenol, p-chloroaniline (PCA), 2,4-dichlorophenol (2,4-DCP), and organic dyes ( Table 4 (17)). In contrast, the persulfate radical (S 2 O •− 8 ) was produced by the reaction of PDS and Mn(III) (Equation (18)). For the system of Mn 2 O 3 /PDS, it is believed that the singlet oxygen ( 1 O 2 ) was the primary active species that was responsible for the degradation of organic pollutants [88]. As demonstrated by Khan [41]. Therefore, the hydroxyl group on the surface of manganese oxides plays a significant role in PDS activation.

Activation of PDS by Mn(III) (Oxyhydr)Oxides
Single or combined Mn(III) (oxyhydr)oxides have been employed to activate PDS to remove different organic pollutants, such as phenol, p-chloroaniline (PCA), 2,4-dichloro phenol (2,4-DCP), and organic dyes ( Table 4). The activation pathway of PDS varies with the different types of Mn(III) (oxyhydr)oxides ( Figure 3). For example, Shabanloo et al reported the generation of active SO 4 • radicals in the nano-Mn3O4/PDS system [87]. Since both Mn(II) and Mn(III) species are identified in the Mn3O4 structure, the formation o SO 4 • was mainly attributed to the activation of PDS by Mn(II) (Equation (17)). In contrast the persulfate radical (S 2 O 8 • ) was produced by the reaction of PDS and Mn(III) (Equation (18)). For the system of Mn2O3/PDS, it is believed that the singlet oxygen ( 1 O 2 ) was th primary active species that was responsible for the degradation of organic pollutants [88] As demonstrated by Khan (20)). The pathway of 1 O 2 formation in the system of A-Mn2O3/PDS is comparable to the approach of producing 1 O 2 in the β-MnO2/PDS system in which the important metastable manganese intermediate was first formed through the complex re action between the hydroxyl group (-OH) and cleavaged S 2 O 8 2 [41]. Therefore, the hy droxyl group on the surface of manganese oxides plays a significant role in PDS activation (17   [89]. The authors reported that the degradation efficiency of phenol in the γ-MnOOH/PDS system was pH-dependent. Under the basic condition (pH 11), phenol was efficiently removed due to the generation of SO •− 4 and HO • radicals. However, at pH 3 and 7, the oxidative intermediate (≡ Mn(III) − − 3 OSOOSO − 3 ) was believed to be responsible for the removal of phenol. Although the mentioned report explained well the oxidation performance of γ-MnOOH/PMS for phenol removal, the information regarding the mechanism of PDS activation on the surface of γ-MnOOH was not given in detail. Considering this, Xu et al. conducted a further investigation focusing on the catalytic mechanism of PDS by γ-MnOOH [90]. Based on the results of chemical scavenging and ESR experiments, a non-radical mechanism was proposed. Generally, the non-radical mechanism in PS activation was attributed to three aspects-the generation of 1 O 2 , the electron transfer process, and the catalyst surface-activated intermediates [91][92][93][94][95]. However, the 1 O 2 production and electron transfer process mechanism were excluded according to the results of ESR and linear sweep voltammetry (LSV) experiments. Therefore, the γ-MnOOH surface-activated PDS molecules were verified as the main active species for the degradation of PCA. Figure 4 shows the formation of active PDS molecules on the surface of γ-MnOOH.
(pH 11), phenol was efficiently removed due to the gen cals. However, at pH 3 and 7, the oxidative intermediate lieved to be responsible for the removal of phenol. Alt plained well the oxidation performance of γ-MnOOH/PM mation regarding the mechanism of PDS activation on th given in detail. Considering this, Xu et al. conducted a f the catalytic mechanism of PDS by γ-MnOOH [90]. Based enging and ESR experiments, a non-radical mechanism w radical mechanism in PS activation was attributed to thre the electron transfer process, and the catalyst surface-However, the 1 O 2 production and electron transfer proce cording to the results of ESR and linear sweep voltammet the γ-MnOOH surface-activated PDS molecules were ve for the degradation of PCA. Figure 4 shows the formatio surface of γ-MnOOH. The activation of PDS by Mn(III) (oxyhydr)oxide co tion was also reported [96][97][98]. For instance, Liu et al Mn3O4 composite (Mn3O4/C) and investigated the reactiv dichlorophenol (2,4-DCP) degradation [96]. The results moval was reached in 140 min and the enhanced degrad ence of the defective edges of the carbon layer, which fac  [97]. The results showed that 40 mg/L of MB was completely removed in 30 min by the system of Ag/Mn 3 O 4 /graphene + PDS under visible light. The enhanced degradation of MB was attributed to the hampered electron-hole recombination due to the loading of Ag and graphene. Furthermore, the studies regarding the application of modified Mn 2 O 3 in oxidants (such as PMS, H 2 O 2 ) activation for contaminants removal were also reported [84,[99][100][101]. For example, Saputra et al. prepared an egg-shaped core/shell α-Mn 2 O 3 @α-MnO 2 catalyst via a hydrothermal process and investigated the catalytic activity of α-Mn 2 O 3 @α-MnO 2 in heterogeneous Oxone®activation for phenol degradation [84].

The Effect of pH
The Mn(III) (oxyhydr)oxides-mediated activation of PDS/PMS can be affected by solution pH in different ways. For example, influencing the property of charge on the surface of the catalysts, changing the ionic forms of PDS/PMS and pollutant molecules, as well as altering the reduction potential of active radicals.
First, the solution pH can affect the interaction between catalyst and PDS/PMS and pollutants through changing the electrostatic effect. The point of zero charges (PZC) of the catalyst and the acid dissociation constant (pKa) of radical precursors and contaminants are two important parameters that are used to recognize the charge type on the surface of the catalysts and the ionic situation of oxidants and pollutants in solution. For instance, when the solution pH is equal to the PZC value of the catalyst, the amounts of positive and negative charges on the surface of the catalyst are equal (i.e., the surface charge of the catalyst is zero). When the solution pH is higher than the PZC value of the catalyst, the surface charges of the catalyst are negative. On the contrary, if the solution pH is lower than the catalyst PZC value, the surface of the catalyst will be positively charged [104]. The same situation is suitable for the analysis of the ionic form of oxidants and pollutants. The PZC values of commonly used Mn(III) (oxyhydr)oxides and the pKa values of PMS/PDS, and some typical pollutants, are summarized in Table 5. The impacts of solution pH on the interaction between Mn(III) (oxyhydr)oxides and PDS/PMS and pollutants have been reported. For example, Zhao et al. reported that the adsorption and degradation of ciprofloxacin (CIP) by the synthesized Mn 3 O 4 -MnO 2 composite were facilitated at neutral pH solution [68]. The results were explained by the enhanced electrostatic attraction between Mn 3 O 4 -MnO 2 and CIP. The PZC value of the Mn 3 O 4 -MnO 2 composite was measured at 2.5; thus, in the solution pH 7, the surface of the catalyst was negatively charged. In comparison, the pKa of CIP was 8.7-10.58, leading to the formation of positively charged CIP ions in the neutral pH solution. Therefore, the electrostatic attraction between the negative catalyst and the positive CIP improved, resulting in a facilitating degradation of CIP. The same phenomenon was also reported in the studies of PDS activation by γ-MnOOH/α-Mn 2 O 3 for pollutant degradation [88,90].
Second, the transformation of radicals also influenced the reactivity of Mn(III) (oxyhydr)oxides for pollutant degradation. For instance, the reported conversion of SO •− 4 to HO • under the basic solution (as shown in (Equation (8)) can have a significant impact. Since the reduction potential value of HO • under natural pH is lower than that in acidic solution (1.8 vs. 2.7V) [105], and the lifetime of HO • is shorter than SO •− 4 (20 ns vs. 30-40 µs) [106]; thus, the transformation from SO •− 4 (E = 2.6 V) to HO • under alkaline solution might lead to a decrease of pollutant degradation. In addition, the leaching of Mn 2+ from Mn(III) (oxyhydr)oxides in an acidic condition also should be taken into consideration for the activation of sulfate compounds (PMS/PDS).

The Effect of Inorganic Anions
Inorganic anions are ubiquitous in various aquatic compartments. It is reported that inorganic anions can suppress the degradation of pollutants in Mn(III) (oxyhydr) oxides-activated PMS/PDS systems through competing with pollutants for radicals. Thus, to evaluate the applicability of the Mn(III) (oxyhydr)oxides + PMS/PDS system in different water matrices, the influence of inorganic anions on the removal of pollutants has been investigated by many researchers [63,79,86,88,97]. In this section, the effect of inorganic anions, such as carbonate/bicarbonate ions (  25)) leading to the inhibited degradation of pollutants [115]. However, although the redox potential of CO •− 3 is low (1.59 V vs. NHE), it can still selectively degrade some organic pollutants with a reaction rate of 10 3 -10 9 M −1 s −1 [116,117]. In addition, the presence of carbonate and bicarbonate ions can affect the stability of oxidants. For example, PDS can be activated by HCO − 3 to generated percarbonate (HCO − 4 ) (Equation (26)) [118]. Similarly, PMS can be catalyzed by both CO 2− 3 and HCO − 3 to form active radicals and HCO − 4 (Equations (27)- (29)). Furthermore, the solution pH can be changed in the presence of carbonate/bicarbonate ions, which can affect the reactivity of Mn(III) (oxyhydr)oxides in PMS/PDS activation as discussed in Section 4.1.
Chloride ion (Cl − ) exists widely in various water bodies including surface water, groundwater, and industrial wastewater [119]. The influence of Cl − on the degradation at Mn-oxides/water interfaces becomes urgent. More experimental work is also needed to develop new Mn-bearing oxides supported with high catalytic efficiency, suitable for industrial applications, and yet are relevant from both economic and environmental points of view.