Current Progress in Advanced Oxidation Processes for the Removal of Contaminants of Emerging Concern Using Peracetic Acid as an Effective Oxidant
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BIOSUV Group, Department of Chemical Engineering, Campus As Lagoas-Marcosende, CINTECX, University of Vigo, 36310 Vigo, Spain
2
Laboratory of Physical Chemistry of the Solid State, Department of Chemical, University of Sfax, Sfax 3000, Tunisia
*
Author to whom correspondence should be addressed.
Catalysts 2025, 15(5), 469; https://doi.org/10.3390/catal15050469 (registering DOI)
Submission received: 24 March 2025
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Revised: 29 April 2025
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Accepted: 8 May 2025
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Published: 10 May 2025
(This article belongs to the Special Issue Advances in Enhancement of Catalytic Performance for Wastewater Treatment)
Abstract
:The growing diversity and prevalence of contaminants of emerging concern (CECs) in aquatic environments present significant risks to human health and ecosystems, necessitating the development of effective remediation strategies. Advanced oxidation processes (AOPs) have emerged as a promising solution due to their ability to produce highly reactive species that efficiently degrade persistent contaminants. Among the various oxidizing agents, peracetic acid (PAA) has attracted significant attention in the field of water treatment for its powerful oxidative properties, environmentally safe decomposition, and ease of use. This article is designed to offer a comprehensive overview of the latest trends in PAA-based AOPs. The discussion begins with an overview of the intrinsic performance of PAA, emphasizing its oxidation potential and degradation mechanisms. Subsequently, the effectiveness of PAA-based AOPs in remediating CECs is explored, focusing on transition metal-mediated activation (Fe, Co, Mn), UV irradiation, and carbon-based catalysts, all of which enhance the generation of reactive species (RS). Next, the determination of RS in PAA-based AOPs is examined, distinguishing between free radical (organic and inorganic) and non-radical (singlet oxygen and high-valent metal) mechanisms that govern pollutant degradation. Then, key factors affecting the removal of CECs in PAA-based AOPs, including initial PAA concentration, catalyst dosage, and pH, are also addressed. Following that, the potential by-products and hazard assessments associated with PAA oxidation are discussed. Finally, current challenges and future research directions are proposed to facilitate the large-scale application of PAA-based AOPs in water remediation.
1. Introduction
The quality and availability of aquatic resources are crucial for sustaining the health of all organisms within ecosystems, ranging from microscopic species to humans [1,2,3]. Access to clean water supports ecosystem stability and enhances human health and well-being. However, in recent times, an increasing presence of unconventional organic pollutants has been detected in natural water bodies, many of which lack regulatory standards, posing significant threats to both ecosystem stability and public health even at environmentally relevant concentrations [4]. The Environmental Protection Agency refers to these pollutants as emerging contaminants; however, since the mid-1990s, the term ‘contaminants of emerging concern’ (CECs) has been commonly used to describe substances such as pesticides, personal care products, industrial chemicals, pharmaceuticals, veterinary antibiotics, and nanomaterials. CECs pose significant risks to human health and aquatic organisms even at low concentrations [5] due to their persistence and bioaccumulation, which is attributed to their high resistance to degradation, a quality linked to their complex structure [6]. Consequently, CEC exposure in humans has been linked to a range of adverse health outcomes, including reproductive system disorders, thyroid dysfunction, Alzheimer’s disease, cancer, and obesity [7]. In wildlife, CECs can also impair reproductive systems, reduce hatchability, and alter vitellogenin levels [7].
Nevertheless, conventional wastewater treatment strategies, including physicochemical and biological methods, often exhibit limited efficiency and may fail to meet the stringent standards imposed by contemporary environmental regulations [8]. Physical methods primarily remove contaminants from water, which may not be adequate for the complete mineralization of pollutants [9]. On the other hand, biological methods may not be suitable for urgent remediation, as they require prolonged periods of use for effective treatment. Furthermore, they are sensitive to changes in pH, temperature, and aeration, all of which are crucial for microbes to efficiently degrade pollutants [10]. These conventional pollutant treatment methods, still widely used for removing most contaminants, exhibit limited efficiency, as they fail to achieve complete destruction. As a result, there is a growing interest in sustainable processes that can better address the challenges posed by these contaminants.
Advanced oxidation processes (AOPs) have emerged as a promising solution for contaminant removal, owing to their environmentally friendly characteristics and wide applicability [11,12]. These processes bridge the gap between traditional physicochemical and biological methods and the current restrictions imposed by environmental legislation [9]. Through the generation of potent reactive radicals, such as sulfate radicals ( [13,14,15], hydroxyl radicals () [16,17,18], and peroxyl radicals ) [19,20,21], AOPs effectively promote the breakdown of bulky organic pollutants into more benign molecules, ultimately leading to their complete mineralization into and [22].
The generation of reactive intermediates is initiated by the activation of oxidants, with the nature of the radicals produced determined by the specific type of oxidant used. In AOPs, commonly used oxidants include hydrogen peroxide [23], ozone [24], persulfate [25], and peroxymonosulfate (PMS) [26]. Since 2019, PAA has gained significant attention as a promising candidate due to its potent oxidative potential, durability, high efficiency at ambient temperature, widespread availability, lower by-product formation, and reasonable cost [27,28,29].
Additionally, its lower O-O bond energy (38 kcal·mol−1) compared to PMS (75.7 kcal·mol−1) and (51 kcal·mol−1) indicates that PAA requires less energy for activation, further enhancing its appeal as an oxidant [30]. This reduced energy requirement facilitates the generation of reactive species (RS) in PAA-based AOPs, which are known to produce and various organic radicals, including acetylperoxyl radicals, acetyloxy radicals , and alkoxyl radicals [19]. Notably, ) radicals are unique to PAA-based AOPs, imparting distinctive characteristics that set them apart from conventional AOPs. Furthermore, PAA not only generates free radicals but also promotes the formation of non-radical species such as singlet oxygen (1) [31], showcasing its versatile oxidizing potential. This versatility positions PAA as a crucial component in advanced water treatment technologies, especially considering the urgent need to manage CECs.
The benefits of PAA make it a compelling candidate for alternative, efficient approaches to removing CECs. In addition, as illustrated in Figure 1, the combination of UV light and PAA can be utilized to achieve simultaneous pathogen inactivation and micropollutant degradation. Furthermore, significant attention has been devoted to understanding the microbial inactivation kinetics of PAA and evaluating how key experimental parameters influence the overall process efficiency.
Therefore, this review aims to explore the role of PAA in AOPs, focusing on its intrinsic performance and activation methods, including metal catalysts, UV irradiation, and carbon-based catalysts. Additionally, we discuss its efficiency in degrading CECs and the determination of the reactive species involved. Furthermore, we examine the various variables influencing its performance. Lastly, we highlight the challenges associated with PAA-based processes and propose potential future avenues for enhancing their applicability in environmental remediation.
2. Intrinsic Performance of PAA
Due to its strong microbicidal effects, PAA has been widely utilized [32]. It is a member of the organic peracid family, characterized by the presence of a -COOOH group. Due to its high redox potential, which ranges from 1.06 to 1.96 V in alkaline and acidic conditions, respectively, PAA has the capability to break down organic micropollutants in water [19,32]. Furthermore, the peroxide group of PAA is thermodynamically unstable. This instability arises from electron repulsion between the oxygen atoms in the peroxide moiety, which further weakens the O-O bond and increases its tendency to decompose, forming hydroxyl and aliphatic radicals. These RS facilitate the degradation of organic pollutants; for this reason, PAA is regarded as an oxidizing agent.
A growing number of recent studies emphasize the use of PAA for treating water contaminated with CECs, underscoring its effectiveness in addressing these challenges. This potency is owed to the distinctive properties of PAA, which functions as a moderately specific oxidant. For instance, Zhang et al. [33] indicated that PAA can directly react with contaminants containing electron-rich, dense moieties, such as unsaturated bonds or conjugated systems. Owing to their high electron density, these groups exhibit greater reactivity toward oxidation, thereby increasing the likelihood of interacting with PAA. As a result, PAA has been shown to effectively oxidize several β-lactam antibiotics, recognized as CECs, including cefalexin, cefadroxil, cefapirin, cephalothin, ampicillin, amoxicillin, and penicillin [33]. Their reactivity with PAA is attributed to their nucleophilic nature, as evidenced by the correlation between the reaction rate constant and the global nucleophilicity of these compounds. Additionally, contaminants containing sulfur moieties exhibit exceptionally high reactivity toward PAA, as the presence of electron-donating groups enhances their interaction with electrophilic species [34]. To improve the understanding of reactions triggered by PAA, Kim et al. [35] published a comprehensive review on the behavior of PAA with the organic compounds of diverse structures and explored the underlying mechanisms. Demonstrating its selectivity, the second-order rate constants for the PAA oxidation of these compounds span nearly 10 orders of magnitude, ranging from to .
However, many organic pollutants, such as ibuprofen, gemfibrozil, and naproxen (NAP), are resistant to degradation at low PAA doses, as PAA is unable to effectively break them down [36]. This limitation is primarily attributed to its low homolytic cleavage rate of , which restricts its effectiveness as an oxidant in technical processes [32]. Moreover, PAA’s high selectivity prevents it from effectively degrading recalcitrant pollutants on its own [36]. This underscores the need for activators to efficiently produce potent oxidizing agents.
3. Proficiency of PAA-Based AOPs in CECs Remediation
In AOPs, PAA has garnered attention as an oxidizing agent capable of producing peroxyl radicals , along with radicals through its activation [37]. The formation of free radicals in PAA-based AOPs was first reported by Caretti and Lubello, who implemented a UV/PAA approach in a wastewater disinfection pilot plant [38]. However, only in recent years have PAA-based AOPs gained recognition as a method for eliminating CECs. Relative to , which is considered the primary oxidant in AOPs, the lower unoccupied molecular orbital (LUMO) energy of PAA (−0.25 eV) indicates its easier activation, suggesting that PAA has a greater capacity to accept electrons than (LUMO energy = 0.57 eV) [39]. Therefore, PAA-based AOPs are regarded as a promising advancement over conventional -based AOPs, with a growing interest in these processes since 2019.
Various methods have been proposed to activate PAA through the cleavage of the O-O bond into two radicals, and (Equation (1)), an important step in the oxidation of micropollutants. Among these, Zhang and Huang proposed a widely accepted mechanism describing the reactions of different radicals in a UV/PAA system, which has since been applied to other PAA-based processes [40]. In addition to degrading organic pollutants, both radicals are also capable of interacting with PAA to produce (Equations (2) and (3)). These radicals can, in turn, undergo bimolecular decomposition, yielding and (Equation (4)). Moreover, the acetoxy radicals formed are also capable of undergoing radical–radical coupling (Equation (5)). On the other hand, to produce methyl radicals (•), undergoes decarboxylation (Equation (6)). Due to their instability, •are reactive intermediates that react immediately upon contact with oxygen to generate peroxyl radicals (Equation (7)), which subsequently undergo bimolecular breakdown (Equations (8)–(10)).
To initiate another oxidation cycle, reacts with again to generate a new PAA molecule, as shown in Equation (11) [40].
Diverse technologies have been applied for PAA activation to facilitate the formation of RS (Figure 1), including the following:
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3.1. Transition Metal for the Activation of PAA
AOPs that involve the activation of PAA by transition metal ions seem highly promising [61]. When transition metals are introduced into a PAA solution, they can easily decompose the O-O bond, leading to the production of and , as evidenced by the following (Equations (12) and (13)). Although PAA activation by transition metal catalysts closely mirrors the traditional Fenton process, it demonstrates greater efficiency in radical generation [62]. This enhanced efficiency is reflected in the reactivity of PAA, which, when combined with , is more than 650 times greater than that of [39]. Additionally, the reaction rate constant for radical generation from PAA is significantly higher (ranging from 0.5 to 1.1 × ) within the pH range of 3 to 7.1, compared to the much lower rate observed for the Fenton reaction involving (K = 63 to 76 ) [39]. This comparison highlights the considerably enhanced reactivity and radical generation efficiency of PAA activation relative to the traditional Fenton process, primarily because PAA decomposition is the main source of radicals in Fe(II)/PAA systems.
Several recent works have investigated the use of transition metals for PAA activation. While the majority of this research has focused on dissolved metal catalysts, some have also explored the use of heterogeneous metal catalysts, such as zero-valent metals and metal oxides. A summary of these findings is presented in Table 1.
3.1.1. Iron-Based Catalysis for PAA Activation
For water treatment, iron-based catalysts have been widely utilized owing to their environmental friendliness, high efficiency, and low cost. Among them, is one of the most extensively used iron catalysts, particularly because of its application in the Fenton process. In systems where PAA is activated by Fe(II)/Fe(III), the pH of the solution plays a critical role, similar to the traditional Fenton reaction. To prevent iron precipitation and maintain catalyst stability, acidic conditions must be carefully maintained [71]. In the first phase, the majority of Fe(II) is consumed to activate PAA and undergo oxidation to Fe(III). At a pH = 3, the performance of contaminant removal is heavily affected by the much lower reaction rate of Fe(III) with PAA (2.72 ) [39], relative to that of Fe(II) with PAA (0.86–1.1 × ) [72]. Therefore, a major challenge for improving Fe(II)/PAA performance is to find efficient strategies to promote Fe(III) reduction back to Fe(II), considering the slow Fe(III)/Fe(II) redox cycle [73].
Several researchers have demonstrated the effectiveness of the Fe(II)/PAA system in removing CECs. For instance, Kim et al. [39] showed that the Fe(II)/PAA system could efficiently eliminate organic dyes, outperforming the conventional Fenton process. Yang et al. [63] reported that the Fe(II)/PAA system rapidly degraded p-arsanilic acid (5 µM), achieving 98% removal within just 20 s using 400 µM of PAA and 200 µM of Fe(II) at a pH of 3.0. In contrast, PAA alone achieved only 15% degradation after 45 min. Lin et al. [74] evaluated the degradation of diclofenac (DCF) by the Fe(II)/PAA system and observed that no significant elimination of DCF (10%) occurred under the conditions tested, which could be attributed to the low dosage of Fe(II) (1 μM) and the slow Fe(III)/Fe(II) redox cycling. To enhance the oxidation capacity of the Fe(II)/PAA process, they added 2,2′-azino-bis(3-ethylbenzothiazoline-6-sulfonate) (ABTS) as a reducing agent, achieving 92% of DCF removal within 30 min. ABTS effectively acted as an electron shuttle, accelerating the Fe(III)/Fe(II) redox cycle.
In recent years, iron-based heterogeneous catalysts have garnered considerable interest in PAA-driven AOPs. For example, Zhang et al. [66] used nano zero-valent iron for the activation of PAA, achieving approximately 95% elimination of tetracycline (TC) within 30 min. Moreover, Wang et al. [62] designed an innovative -modified zeolite/PAA system for the removal of sulfamethoxazole (SMX) at a neutral pH, achieving complete degradation within 60 min. Furthermore, Wang et al. [52] investigated the performance of the dual-metal catalyst, cobalt ferrite (), combined with the PAA system for the removal of SMX as a target pollutant. The system achieved 87.3% elimination within 30 min under neutral conditions, using 200 μM of PAA and 0.1 g·L−1 of . Furthermore, Wang et al. [75], applied pyrite tailings (PRT) to activate PAA for the removal of SMX. In situ Raman spectra confirmed the activation of PAA by PRT, with 93.67% of SMX (5 mg·L−1) removed within 50 min under near-neutral conditions (pH of 5.8). In this system, PAA activation was primarily attributed to leached Fe ions from PRT. Spectroscopic analyses and theoretical calculations revealed that sulfur sites in PRT enhanced Fe(II) regeneration, thus promoting the degradation process. Radical quenching experiments confirmed that acetylperoxyl and hydroxyl radicals were responsible for SMX degradation.
3.1.2. Cobalt-Based Catalysis for PAA Activation
Cobalt (Co) stands out as a highly efficient catalyst for radical generation, outperforming metals such as manganese, copper, and iron in PAA activation. This superior catalytic activity is attributed to the high standard redox potential of ( = 1.82 V) and the ability of PAA to rapidly reduce to without external energy input [76].
The decomposition of PAA by Co leads to the formation of and as the primary reactive radical species without the formation of . This process is illustrated by the reactions presented in Equations (14) and (15), which detail the generation of these reactive radicals [77].
Kim et al. [20] found that the apparent second-order reaction rate constants for Co(II) and Co(III) with PAA (1.70 × –6.67 × and 3.91 × –4.57 × ), respectively, were comparable, based on kinetic simulations conducted over an initial pH range of 3.0–8.2. The comparable reaction rates of PAA with Co(II) and Co(III), along with the sustained Co(II)/Co(III) cycling, allow for the use of a low Co dosage in this AOP, improving the feasibility of this technology while minimizing concerns about Co toxicity. Several studies have examined the effectiveness of the /PAA process for various contaminants. For instance, Wang et al. [77] achieved approximately 90% degradation of SMX in 15 min through PAA activation using a minimal amount of Co (0.8 μM). The use of scavengers led to the conclusion that the organic radicals and played the primary role in degradation. Building on these findings, they proposed a reaction pathway outlining the involvement of RS in the Co/PAA system for eliminating the pollutant, as illustrated in Figure 2. Using the Co/PAA system, Kim et al. [20] conducted a comprehensive study on the elimination of 30 aromatic organic compounds, confirming as the primary active free radical.
On the other hand, a recent investigation by Liu et al. [41] revealed that under acidic conditions, the /PAA system can generate species, which may contribute to the degradation of organic contaminants. The pathway proposed for the formation of species is illustrated in Equations (16) and (17).
However, unlike other metal/PAA systems, one of the advantages of /PAA is its ability to function under neutral conditions, which has led to the previously overlooked role of species. Neutral pH appears to favor the formation of oxygen-centered radicals, such as and . Despite this, it warrants consideration, as certain pollutants, such as carbamazepine (CBZ), exhibit greater reactivity with than with oxygen-centered radicals [41].
Moreover, the Co(II)/PAA process faces challenges, such as limited reusability and risks of secondary pollution, which further restrict its applicability in PAA-based AOPs. In addition to these limitations, the potential ecotoxicological risks associated with Co leaching must also be carefully considered. The leaching of Co ions into the environment can have harmful effects on both aquatic and terrestrial ecosystems [78]. The toxicity of Co is highly influenced by its oxidation state and solubility, as well as various intrinsic and extrinsic factors [79]. Excessive exposure to Co is harmful to humans, potentially causing various diseases such as contact dermatitis, pneumonia, allergic asthma, and lung cancer [80], which limits its direct use in a homogeneous form.
Therefore, further research is needed to investigate the catalytic performance and mechanisms of cobalt-based heterogeneous systems. To overcome these drawbacks, cobalt-based catalysts can be engineered into heterogeneous forms, such as [69] and [68], which have been designed to catalyze PAA. Among these, particles exhibit outstanding catalytic performance, which is attributed to the coexistence of Co(II) and Co(III) within their structure. Nevertheless, certain pollutants, such as DCF, benzoic acid (BA), and CBZ, exhibit significant recalcitrance and greater resistance to the process [68]. In addition, zero-valent cobalt (ZVCo) has been developed as an excellent heterogeneous catalyst for PAA, targeting the removal of SMX as a contaminant. Zhou et al. [30] demonstrated that 99% SMX was removed by the ZVCo/PAA system within 5 min under neutral conditions. Additionally, ZVCo can be easily recycled due to its magnetic properties, which enhance its practicality and reusability in the system. Furthermore, even after four cycles, ZVCo maintained its remarkable ability to activate PAA for SMX degradation, highlighting its stability and reusability.
Due to its effectiveness in PAA activation, Co appears to be a promising option for the degradation of CECs. However, it is important to recognize that Co is associated with potential risks to human health. Therefore, Co should be neutralized before discharge in processes that utilize homogeneous Co, while heterogeneous systems must account for the potential leaching of Co into the sample. To mitigate these risks, several strategies can be implemented to minimize cobalt leaching in Co/PAA systems. These strategies aim to reduce Co release into the environment, enhance the efficiency of the catalytic process, and promote sustainable practices. Among these, cobalt-doped graphitic carbon nitride has shown promise in stabilizing ions, thereby preventing their release during catalytic reactions and improving the catalyst’s efficiency in pollution degradation [81]. Similarly, nano-island-encapsulated cobalt single-atom catalysts provide enhanced cobalt stability by embedding single cobalt atoms within a support matrix, further minimizing leaching while optimizing catalytic performance [82]. Additionally, chitosan-derived carbon has been explored to reduce cobalt leaching in disinfectant degradation, as its carbon matrix efficiently encapsulates cobalt-based nanoparticles, preventing metal release and maintaining catalytic efficiency [83]. These strategies contribute to improving the overall environmental impact of cobalt-based systems.
3.1.3. Manganese-Based Catalysis for PAA Activation
Manganese (Mn), a widely abundant transition metal, exhibits a broad range of oxidation states, from +2 to +7, in aqueous solutions and offers greater environmental friendliness compared to Co [84]. Unlike iron, it functions effectively across a wide pH range without generating sludge formation [85]. Mn(II) has been extensively studied as a homogeneous catalyst for activating various oxidants [86,87,88]. However, the role of soluble Mn(II) in mediating PAA activation remains insufficiently explored, likely due to its relatively low catalytic efficiency [89]. Consequently, there is limited information on the broader use of the Mn(II)/PAA system for eliminating organic pollutants. The only reported instance involves the successful degradation of the organic dye Orange II using Mn(II) salt in combination with PAA [43]. This study demonstrated that Orange II was degraded at a pH of 9.4; however, a high dosage of PAA ([PAA] = 5–20 mM, [Mn(II)] = 100 μM) was required. In contrast, degradation at acidic to neutral pH levels (pH 3–7) was significantly less effective under similar conditions. Further analysis revealed that Mn(IV), Mn(VI), and Mn(VII) were the primary reactive species generated during the reaction, as confirmed by UV–visible spectral changes and electron paramagnetic resonance (EPR) spectroscopy [43].
The limited performance of the Mn(II)/PAA system may stem from the inherent instability of high-valent Mn species, which are prone to spontaneous decomposition or disproportionation in water [90]. To overcome this challenge, the use of ligands in transition metal-based AOPs has been widely reported as an effective strategy for enhancing the degradation efficiency of organic pollutants [89,90,91]. By coordinating with transition metal ions, ligands form stable metal–ligand complexes that improve system stability and overall performance. Building on this principle, recent research has focused on the introduction of ligands. For example, Liu et al. [91] demonstrated the improved degradation of SMX by Mn(II)-activated PAA with the addition of nitrilotriacetic acid (NTA), which complexes with Mn(II) to form the Mn(II)-NTA complex, accelerating PAA activation. In the Mn(II)/PAA/NTA process, Mn(V) played a major role in SMX elimination. Notably, the Mn(II)/PAA/NTA system exhibited high treatment efficiency under acidic and neutral conditions, emphasizing its potential for wider environmental applications. Dong et al. [90] further found that the addition of ethylenediamine-N,N′-disuccinic acid (EDDS) significantly enhanced the elimination of atrazine (ATZ) in the Mn(II)/PAA process. Moreover, in the Mn(II)/EDDS/PAA system, Mn(V) was confirmed as the primary reactive species responsible for ATZ removal, as the inclusion of the ligand improved the stability of manganese intermediates. Consequently, the role of high-valent metals (HVMs)—transition metals in oxidation states higher than +3, capable of facilitating strong oxidative transformations—in transition metal-catalyzed PAA systems is of considerable importance and cannot be overlooked. For instance, Kim et al. [89] reported that Mn(II) alone exhibits limited reactivity with PAA; however, the addition of picolinic acid (PICA) significantly enhances the decomposition of PAA by Mn(II). The PAA-Mn(II)-PICA system efficiently removes various contaminants, including bisphenol A (BPA), NAP, SMX, CBZ, and trimethoprim, at neutral pH, achieving over 60% removal within 10 min in both pure water and wastewater environments.
However, the incorporation of organic ligands to enhance the Mn(II)/PAA process may elevate chemical oxygen demand levels or lead to the formation of by-products through reactions with disinfectants. Moreover, their use adds chemical complexity to the wastewater, requiring additional removal steps. To improve the oxidation potential of the Mn(II)/PAA process, it is crucial to investigate more economical and eco-friendly approaches.
To improve the oxidation potential of the Mn(II)/PAA process, it is crucial to investigate more economical and eco-friendly approaches. Several strategies can be evaluated, including the use of green ligands to mitigate the environmental burden on wastewater and the exploration of the feasibility of recovering organic ligands post-use.
3.2. UV Irradiation-Induced Activation of PAA
The activation of PAA by UV irradiation is among the most extensively studied processes, as UV radiation initially induces the homolytic cleavage of the O-O bond in PAA, generating and (Equation (18)) [40,92,93]. For this purpose, the most commonly used UV light wavelength is 254 nm, which is emitted by low-pressure mercury lamps. Zhang et al. [40] quantified the quantum yield for PAA decomposition at a UV of 254 (Φ = 0.88 ± 0.04 mol-Ein−1), which exceeds that of (Φ = 0.5 mol−1), indicating a faster rate with a comparable amount of absorbed light. Since always presents in equilibrium with PAA solutions, it also participates in the degradation processes under ultraviolet radiation, as shown in Equation (19).
For the degradation of CECs, numerous recent scientific publications have focused on utilizing the UV/PAA process, and these findings are summarized in Table 2.
Notably, UV light alone can directly degrade certain organic contaminants that have dissociative absorption bands in the UV range [94]. However, in most cases, achieving the direct UV mineralization of organic pollutants is challenging. Nevertheless, when PAA is combined with UV irradiation, the CECs can be destroyed by various mechanisms, offering a potential solution to this challenge: first, direct oxidative degradation, utilizing the strong oxidizing property of PAA; second, direct photolysis, where organic pollutants absorb photons from ultraviolet light and undergo degradation; and third, the well-known activation of PAA through UV irradiation, leading to the production of highly active radicals such as and . The presence of in the UV/PAA system is thought to be responsible for its greater selectivity compared to the UV/ process [53]. A number of studies have shown that the UV/PAA system can improve the efficiency of pollutant degradation through a synergistic effect, suggesting that UV/PAA is a promising technology for water treatment. For instance, Sharma et al. [95] reported degradation rates ranging from 80% to 100% for various chlorophenols. Their findings indicate that the position and number of chlorine atoms in the aromatic ring structure play a significant role in the process. Similarly, Cai et al. [21] investigated the effectiveness of the UV/PAA process in degrading several pharmaceutical compounds. They demonstrated that, at a pH of 7.1, while PAA alone resulted in minimal degradation, all compounds were degraded by more than 93.5% with 1 mg·L−1 of PAA under UV irradiation, further highlighting the effectiveness of the UV/PAA process.
To identify the reactive species responsible for the degradation, they also conducted scavenging experiments. They stated that the removal was primarily driven by oxidation, succeeded by oxidation by other radicals, mainly direct photolysis and organic radicals [40]. Additionally, organic radicals reacted rapidly with NAP but not with ibuprofen or carbamazepine, demonstrating their greater selectivity and suggesting a higher affinity for specific pollutants [40].
Multiple factors, such as PAA dosage and pH, have been identified as influencing the performance of the UV/PAA system. Among these factors, UV exposure at 254 nm plays a crucial synergistic role, as it is a light-activated system, as demonstrated by multiple studies [21,40]. On the other hand, radical generation in solution is constrained by the system’s photon absorption capacity [93]. Nonetheless, based on the available literature, the impact of the reactor design and configuration on the UV/PAA system remains largely unexplored.
Table 2.
A summary of studies using UV irradiation for PAA activation in the removal of CECs.
| CEC | [CEC]0 | Removal Efficiency (%) | Activator | PAA Dose | pH | Reaction Time | Ref. |
|---|---|---|---|---|---|---|---|
| Fluoxetine | 5 mg·L−1 | 100 | UV (254 nm) irradiation: (647–3502) W·m−3 | 5–100 mg·L−1 | 7 | 30 min | [53] |
| Sulfamethoxazole | 5 mg·L−1 | 100 | UV (254 nm) irradiation: (647–3502) W·m−3 | 5–100 mg·L−1 | 7 | 30 min | [53] |
| Diclofenac | 1 μM | 90 | UV (254 nm) intensity: 2.12 × 10−6 E·L−1·s−1 | 1 mg·L−1 | 7.1 | <5 min | [21] |
| Ibuprofen | 1 μM | 90 | UV (254 nm) intensity: 2.12 × 10−6 E·L−1·s−1 | 1 mg·L−1 | 7.1 | 30 min | [21] |
| Carbamazepine | 1 μM | >90 | UV (254 nm) intensity: 2.12 × 10−6 E·L−1·s−1 | 1 mg·L−1 | 7.1 | 30 min | [21] |
| Naproxen | 1 μM | >95 | UV (254 nm) intensity: 2.12 × 10−6 E·L−1·s−1 | 1 mg·L−1 | 7.1 | 10 min | [21] |
| Chloramphenicol | 25 mg·L−1 | 100 | UV (254 nm) doses: 0~12.5 W·m−2 | 5–50 mg·L−1 | Not mentioned | 120 min | [96] |
| Clofibric acid | 1 μM | >90 | UV (254 nm) intensity: 2.12 × 10−6 E·L−1·s−1 | 1 mg·L−1 | 7.1 | 10 min | [21] |
| Venlafaxine | 5 mg·L−1 | 100 | UV (254 nm) irradiation: (647–3502) W·m−3 | 5–100 mg·L−1 | 7 | 30 min | [53] |
| 4-chlorophenol | 4 μM | >90 | UV (254 nm) | 3042 mg·L−1 | 9.5 | 5 min | [54] |
| Pentachlorophenol | 4 μM | 100 | UV (254 nm) | 3042 mg·L−1 | 9.5 | 5 min | [54] |
| 2,4,6-trichlorophenol | 4 μM | >80 | UV (254 nm) | 3042 mg·L−1 | 9.5 | 5 min | [54] |
| 2,4-dichlorophenol | 4 μM | 100 | UV (254 nm) | 3042 mg·L−1 | 9.5 | 5 min | [54] |
| Bezafibrate | 1 μM | 80 | UV (254 nm) intensity: 2.12 × 10−6 E·L−1·s−1 | 1 mg·L−1 | 7.1 | 120 min | [21] |
| Carbamazepine | 5 mg·L−1 | 100 | UV (254 nm) irradiation: (647–3502) W·m−3 | 5–100 mg·L−1 | 7 | 30 min | [53] |
| Naproxen | 4 μM | 80 | UV (254 nm) intensity: 9.04 × 10−8 E·L−1·s−1 | 20 mg·L−1 | 7 | 14 min | [92] |
3.3. Carbon-Based Catalyst for the Activation of PAA
Green, metal-free catalysts mark significant advancements in modern chemistry, with their widespread adoption offering great promise due to their ability to completely eliminate toxic metal leaching and secondary water contamination [59]. Among these, carbon materials stand out with remarkable advantages in terms of stability, activity, and renewability in comparison to traditional metal or metal oxide catalysts. Additionally, their cost-effectiveness and environmental friendliness position them as compelling alternatives to conventional catalysts, aligning with the goals of sustainable development [60].
Carbon materials occur in various allotropic forms, including activated carbon [59], carbon nanotubes [97], biochar [60], and graphene [98]. In the field of water treatment, carbon materials not only serve as adsorbents with excellent capacity but also function as catalysts to enhance PAA activation for the removal of organic pollutants. Table 3 presents selected studies employing carbon materials for PAA activation.
This capability was first demonstrated in 2015 by Zhou et al. [99], who combined activated carbon fibers (ACFs) with PAA for the removal of the dye Reactive Brilliant Red X-3B, achieving nearly 97% removal after 45 min. Similarly, in 2022, another example concerned the use of carbon nanotubes (CNTs), where pristine CNTs were utilized to activate PAA and characterize the reactive species and associated active sites within the CNTs [97].
The dominant species responsible for the elimination of BPA was the CNT-PAA* complex, while a secondary role was played by adsorbed radicals. Additionally, the adsorption energy of PAA on the CNT edge was stronger than that on the CNT matrix, as shown by density functional theory calculations, suggesting a favorable interaction between CNT edges and PAA, which facilitates PAA activation for adsorbed generation.
Furthermore, biochar is noted for its durability, insolubility, and high carbon content [100]. Its surface contains numerous defect structures and oxygen-containing functional groups, which are capable of donating electrons to oxidants and producing free radicals [101]. In a study by Hou et al. [102], they investigated the activation of PAA using biochar derived from chicken manure for the degradation of SMZ. Their study demonstrated that was identified as the dominant reactive substance, which was produced via electron transfer. In another study, Zhang et al. [103] explored the oxidation of SMZ using PAA activated by biochar, which was thermally modified under neutral environmental conditions. Their EPR and free radical quenching experiments clearly demonstrated the existence of 1 in the system, which plays a dominant role in the degradation of SMZ. One such material, graphene, is composed of a two-dimensional framework, formed by the organization of carbon atoms in a honeycomb pattern via sp2 bonding [104]. Due to its unique structure, graphene exhibits hydrophobic characteristics, which pose challenges for its application in AOPs. However, researchers have made considerable advancements by investigating graphene-based derivatives, such as graphene oxide, reduced graphene oxide, and other more complex synthetic materials. These derivatives exhibit their adaptability in AOPs [105], particularly when oxygen-containing functional groups, such as hydroxyl, carboxyl, and carbonyl groups, are present. These groups reduce the hydrophobicity of GO, significantly enhancing its reactivity in AOPs [106]. Additionally, GO can promote mediated electron transfer through its intercalated structure, further improving catalytic removal efficiency in AOPs [107]. In this context, Tshangana et al. [108] utilized the graphene oxide quantum dots/PAA system for the removal of sulfasalazine, demonstrating its effective degradation. To reduce the insulating property of GO, rGO is engineered from it [109]. As a result, rGO displays a greater density of defects and a notably increased specific surface area in comparison to the original material [110]. Furthermore, rGO facilitates the production of stable intermediates by activating oxidizing agents due to its distinctive single-layer structural arrangement [105].
Table 3.
A summary of studies employing carbon materials for PAA activation in CECs removal.
| CEC | Catalyst | Processes | [CEC]0 | Removal Efficiency (%) | PAA Dose | PH | Reaction Time | Ref. |
|---|---|---|---|---|---|---|---|---|
| Reactive Brilliant Red X-3B | ACFs | ACFs/PAA | 50 μM | 97 | 5 mM | 7 | 45 min | [99] |
| sulfamethoxazole | rGO | rGO/PAA | 10 μM | 100 | 0.1 mM | 5 | 5 min | [98] |
| bisphenol A | CNTs | CNT/PAA | 0.02 mM | 96.4 | 0.25 mM | 7 | 20 min | [97] |
| phenol | CNT950 | CNT950/PAA | 20 μM | 100 | 0.15 mM | 10 | 10 min | [111] |
| sulfamethoxazole | AC600 | AC600/PAA | 20 mg·L−1 | 99.4 | 0.26 mM | 7 | 150 min | [59] |
| sulfamethazine | activated biochar (ABC) | ABC/PAA | 5 mg·L−1 | 72.8 | 0.07 mM | 7 | 100 min | [103] |
| phenol | carbonized polyaniline (CPANI) | CPANI/PAA | 10 μM | 97 | 0.1 mM | 7 | 60 min | [112] |
The degradation mechanisms of the carbon material/PAA systems are currently classified into three main types:
- -
- The first mechanism involves the generation of singlet oxygen 1, generated through PAA activation by N-doped carbonaceous catalysts. For example, Tian et al. [112] employed carbonized polyaniline (CPANI) as the catalyst, where the C=O group on CPANI activated PAA to generate 1, the main reactive species driving phenol degradation.
- -
- The second mechanism is based on direct electron transfer (DET). In this context, Kong et al. [98] demonstrated the effectiveness of reduced graphene oxide (rGO) in activating PAA for the rapid removal of SMX, achieving a near-complete elimination of the pollutant within just 2 min. Through a combination of quenching experiments, open-circuit potential measurements, and probe-based studies, they confirmed that DET was the dominant degradation pathway. The rGO/PAA system exhibited strong removal performance even in complex water matrices, highlighting the advantages of DET-driven oxidation. Building upon these findings, subsequent research explored other carbon-based materials to enhance DET processes. For instance, Kong et al. [111] emphasized the role of the physicochemical properties of CNTs on organic pollutant removal and PAA activation. The CNT/PAA system was also dominated by the DET oxidation pathway, achieving high MP removal rates. The enhanced catalytic efficiency of surface-regulated CNTs was attributed to reinforced DET, driven by the increased oxidative potential of the CNT/PAA complex and the enhanced electrical conductivity of CNTs. Furthermore, the larger specific surface area and lower oxygen content of CNTs were found to contribute significantly to the elevated oxidative potential, with electrical conductivity closely linked to their degree of graphitization.
- -
- The third degradation mechanism involves the participation of active radicals, as demonstrated by a study on the degradation of 4-chlorophenol (4-CP) via PAA activation using a newly synthesized biochar (P5S5-SDBC) [60]. Mediated electrochemical oxidation and reduction tests revealed that persistent free radicals (PFRs), generated by structural defects in the biochar, acted as electron shuttles, enhancing PAA activation and promoting the formation of reactive species. Electron paramagnetic resonance (EPR) measurements further confirmed that radicals were the primary active species responsible for 4-CP degradation [60].
In recent years, PAA-AOPs, particularly those utilizing carbon-derived catalysts, have attracted considerable attention in environmental remediation. However, carbon materials often suffer from limited stability, primarily due to their vulnerable edges, surface defects, and susceptibility to interactions with reaction intermediates through π–π stacking or hydrogen bonding [113]. To address these challenges, the strategy of encapsulating metal cores within robust carbon shells has emerged as a promising solution. This approach effectively protects the internal metal centers, enhances catalyst stability, minimizes metal leaching, and reduces the risk of secondary pollution. Despite these advances, research into the use of metal–carbon composite materials for contaminant removal via PAA activation remains relatively limited and warrants further exploration.
3.4. Comparative Analysis of Catalytic System for PAA Activation: Transition Metals, UV Irradiation, and Carbon-Based Catalysts
Various catalytic systems, such as transition metals, UV irradiation, and carbon-based catalysts, have been explored to improve the activation of PAA. Transition metal catalysts are widely utilized due to their ability to generate ROS, including and . They are cost-effective and abundant, making them an attractive option for various applications. Furthermore, systems using transition metals are straightforward to implement, simplifying operational procedures. However, their long-term stability and activity may decrease over time due to factors like oxidation, which limits the overall effectiveness and longevity of the catalyst. The use of transition metals in catalytic processes can lead to secondary pollution due to the leaching of metal residues. This may result in the accumulation of harmful metal ions in treated water, posing environmental risks. In homogeneous catalytic systems, it is challenging to recover transition metal catalysts once they have been used in the reaction. This complicates the recycling process and increases the overall operational costs. In contrast, the UV irradiation activation of PAA does not require any chemical additives, making it an environmentally friendly option that minimizes the risk of by-product generation. The simplicity of operating UV systems adds to their appeal. However, UV activation is energy-intensive, which can increase operational costs. Additionally, UV may struggle to penetrate turbid or colored water, thereby reducing its effectiveness. Carbon-based catalysts are known for their environmental compatibility, as they are non-toxic and eco-friendly. They offer a high surface area for both catalytic reactions and adsorption, which enhances their performance. Another advantage is their reusability, contributing to sustainability in long-term applications. On the downside, carbon-based catalysts are typically less efficient in activating PAA compared to metal-based catalysts. The cost of producing high-quality carbon materials can be prohibitive, and some carbon materials require modifications to increase their reactivity for optimal performance.
Each catalytic system presents distinct advantages and limitations. Selecting the appropriate process depends on specific treatment objectives, environmental considerations, and economic factors.
The proficiency of PAA-based AOPs in the remediation of CECs is closely linked to the nature of the RS involved. PAA, upon activation, generates highly RS, such as , , 1, and HVMs, all of which play pivotal roles in the oxidative elimination of CECs. The determination of these RS is essential for understanding the underlying mechanisms of PAA-based AOPs, as their generation directly influences the efficiency and selectivity of pollutant degradation. Through techniques such as scavenging experiments, isotope labeling, and spectroscopy, researchers can identify and quantify these RS, providing crucial insights into optimizing operational conditions and improving the overall performance of PAA-based AOPs in environmental remediation.
4. Determination of Reactive Species in PAA-Based AOPs for Water Decontamination
The efficiency of CEC treatment predominantly depends on the RS generated during the PAA activation process. These include free radicals such as [46], , , and • [50,66], as well as non-radical species like 1 [103], HVMs like Fe(IV) and Co(IV) [114] (Figure 3), electron-transfer processes [98], and direct oxidation [36].
To provide deeper insights into the degradation mechanisms of organic pollutants in PAA-based AOPs, it is crucial to identify the primary RS driving the degradation process. Based on prior research, • has been identified as a less significant radical in pollutant degradation due to its limited oxidizing ability and its propensity to react with dissolved oxygen, forming [19]. Notably, the oxidizing strength of is comparable to that of •, further diminishing its effectiveness in degrading pollutants.
Additionally, the spontaneous decomposition rate constants of , as determined via Equations (5) and (6), are and , respectively. It is worth noting that its spontaneous decomposition rate typically exceeds its reaction rate with pollutants [66], which may limit its direct contribution to degradation. However, this behavior aligns with previous studies indicating that , along with other oxygen-centered radicals, such as , play crucial roles in oxidative degradation pathways [20,21,35,40,115].
4.1. Free Radicals
4.1.1. Organic Radicals: and
Prior studies have demonstrated that the C=C double bond is highly prone to attack by organic radicals. For instance, the reaction rate constant between and the C=C double bond in β-carotene ) is , indicating that β-carotene can effectively neutralize [116]. Similarly, 2,4-hexadiene (2,4-HD), which also contains a C=C double bond, is widely recognized for its ability to scavenge , with a rate constant of approximately for the reaction. Furthermore, due to its fast reaction with (k2,4-HD,•OH = 9.2 × 109 ), (2,4-HD) can simultaneously scavenge both and . Consequently, the inhibitory effect of tert-butyl alcohol (TBA), which is typically a scavenger of and 2,4-HD, on compound degradation may vary, and this variation can demonstrate the role of in the system [59,66]. In this context, Zhang et al. [117] revealed that can act as a specific scavenger for in PAA-based AOPs via electron transfer, without significantly influencing the decomposition of PAA itself [117].
The core concept behind using the quenching method to identify a specific active species is that the quencher must not interfere with the interactions between pollutants and other compounds within the system. For instance, in the CNT/PAA system, introducing 2,4-HD markedly reduces the adsorption of BPA by CNTs [97]. Even at minimal concentrations, 2,4-HD induces the aggregation of CNT powder in the solution, thereby affecting the overall reaction conditions of the system. To mitigate these effects, bromotrichloromethane emerges as a novel alternative, effectively terminating without altering the catalyst’s adsorption properties, making it the most suitable scavenger available [97,118].
On the other hand, for the identification of organic radicals, other methods such as the probe technique can be used. This approach relies on introducing trace amounts of probe compounds, allowing chemical probe experiments to determine the type and presence of active species based on the degradation of the probe molecules and the characteristics of the resulting products. For example, 2,2,6,6-tetramethyl-1-piperidinyloxy (TEMPO) is a stable nitroxide radical that acts as a radical probe for detecting and . The generated methyl-TEMPO (-TEMPO) can be analyzed using mass spectrometry [52]. For instance, during the pyrite/PAA process, the generation of -TEMPO is evidenced by the -TEMPO signal (m/z 216.1598 and a retention time of 1.239 min) [71].
Additionally, the EPR method is frequently utilized to accurately detect the presence of . 5-diisopropoxyphosphoryl-5-methyl-1-pyrroline N-oxide (DIPPMPO) serves as a widely accepted spin-trapping agent. The extracted hyperfine splitting constants (aP, aN, and aH) of the captured radicals are then simulated using Easyspin software (version 5.2.28 toolbox for MATLAB (The MathWorks, MA, USA)), which also aids in analyzing the resulting fitting curve. This process confirms the presence of , •, and [119,120]. However, the detection of remains inconclusive, likely due to its high instability and the absence of relevant fitting information [98]. Recently, Chen et al. [121] conducted a comprehensive study on radical identification through in situ EPR analysis, successfully establishing a reliable method for detecting in PAA-based AOPs by trapping it with DMPO in ethanol under a nitrogen atmosphere.
4.1.2. Inorganic Radicals:
In the context of PAA-based AOPs, represents the principal inorganic reactive species, noted for its exceptionally high reactivity, with reaction rate constants for different compounds typically in the range of [122]. Within such oxidative environments, alcohols are widely acknowledged as potent radical scavengers. For instance, methanol (MeOH), ethanol (EtOH), and tert-butyl alcohol (TBA) effectively scavenge , preventing radical-induced reactions, with respective rate constants of , (1.2–2.8) , and (3.8–7.6) [122,123]. Nevertheless, MeOH and EtOH are prone to reacting with organic radicals, subsequently generating aldehydes or alcohol-derived radicals [55]. Conversely, in PAA-based AOP systems, TBA uniquely acts as a highly selective scavenger for, as clearly depicted in Figure 4. The substantial decline in pollutant degradation following the introduction of TBA strongly underscores the indispensable contribution of within these oxidative systems [21,55].
In addition to alcohols, several other compounds can be effectively utilized to detect the presence of radicals in AOPs. For example, nitrobenzene (NB) and p-chlorobenzoic acid (p-CBA) rapidly react with , exhibiting reaction rate constants of 3.9 and 5.0 , respectively [55,77]. Additionally, these compounds demonstrate negligible reactivity toward other reactive species, rendering them highly selective and dependable probes for detection. Consequently, the measurable reduction in the concentrations of these probe compounds serves as strong evidence confirming the presence and activity of .
On the other hand, in certain heterogeneous systems, commonly employed alcohols may not completely quench surface-bond due to substantial electrostatic repulsion exerted by the catalyst surface. Given its high reactivity, reacts with potassium iodide (KI) at a rate constant of 1.2 [97,124], and KI has been considered as an alternative probe for detecting surface-bound in such a system. Moreover, sodium azide () is also recognized as an efficient quencher of surface-bound radicals, with a reaction rate constant of 1.2 for [125]. Notably, its low dielectric constant (4.5), compared to EtOH (25.3) and TBA (12.5), enhances its suitability for quenching surface-bound OH in heterogeneous systems [126]. It is relevant to point out that KI and act as powerful electron donors, capable of directly reacting with PAA, leading to its dissipation and subsequently hindering the elimination of model compounds [127]. Therefore, when using KI or as quenchers, the careful monitoring of changes in PAA concentration is essential. To offset potential depletion, the supplemental addition of PAA into the system is recommended during quenching experiments.
4.2. Non-Radical Species
4.2.1. Singlet Oxygen (1)
1 is a highly selective oxidant compared to other inorganic radicals. It can react with unsaturated organic compounds through electrophilic attack or by abstracting electrons [122]. 1 exhibits a lower redox potential (0.81 V) compared to , indicating that its reaction rate with most organic compounds is generally slower than that of radicals [128]. The reaction rate of 1 with organic compounds typically ranges from to . According to earlier research, sites that favor the production of 1 comprise graphitic N, electron-deficient carbon, N-metal sites, oxygen vacancies, and ketone groups [129,130]. In PAA-based AOPs, the dissociated PAA anion can react with the central carbon atom of PAA to produce 1, as shown in Equation (20) [131].
A widely used scavenger for 1 is furfuryl alcohol (FFA), which is acknowledged as a highly effective quencher. The rate constant for the reaction between FFA and 1 in aqueous solutions is [59,132]. The steady-state concentration of 1 is calculated using the pseudo-first-order degradation rate constant with FFA as the target contaminant (Equation (21)) [133].
On the other hand, FFA reacts effectively with , with a reaction rate constant of . Therefore, excess TBA must be added during the quenching process if is generated [59]. Wang et al. [55] reported that CBZ reacts rapidly with 1 as well as . To scavenge , excess TBA is used, and the removal of CBZ serves as a clear indicator of the existence of 1. Additionally, the reaction rate of β-carotene with 1 ) highlights its role as an effective 1 scavenger [134].
In EPR measurements, 2,2,6,6-tetramethyl-4-piperidinol (TEMP) is commonly selected as a trapping agent for 1. Its oxidation by 1 leads to the formation of TEMPO, which produces a distinct triplet signal. For instance, TEMPO can be detected in the thermoactivated PAA process, with its production rate increasing as the temperature elevates [55].
Moreover, prior studies have demonstrated that the lifetime of 1 in heavy water (deuterium oxide, ) ranges from 20 to 32 μs, which is more than tenfold longer than in (approximately 2 μs), thereby extending the reaction window. This prolonged lifetime enhances the degradation efficiency of pollutants mediated by 1. The increased pollutant removal observed in further underscores the significant role of 1 in oxidative processes [135].
Furthermore, numerous studies have reported that 1 serves as a major reactive species for degrading organic compounds in solvents [136,137]. However, these conclusions require stronger validation through additional experiments. Notably, water itself can act as a scavenger of 1, with a reaction rate constant of . Since water typically serves as the medium and exists in concentrations much higher than those of organic pollutants in aqueous environments, it is hypothesized that 1 contributes significantly to the degradation of organic compounds. For this to occur, the reaction rate between 1 and pollutants must exceed , based on competitive kinetics.
Given these considerations, is recommended for use in PAA-based AOPs to leverage its ability to extend the lifetime and effectiveness of 1.
4.2.2. High-Valent Metals (HVMs)
HVMs generated during PAA activation have recently been recognized as key intermediates [130]. Unlike radicals, HVMs exhibit longer lifetimes, potentially making them more stable and effective under certain conditions [138]. In contrast, radical-mediated oxidation processes are often impeded by the presence of natural organic matter (NOM), including oxyanions and halide ions, which are commonly found in water matrices and significantly reduce their efficiency [139,140]. Similarly, 1 exhibits substrate-specific reactivity [141] and relatively low oxidation capability [128]. This has led to significant debate regarding whether 1 serves as the primary reactive oxygen species responsible for pollutant degradation during AOPs [122]. Conversely, HVMs exhibit strong oxidation potential, as indicated by [142], [143], and [144]. Their effectiveness in contaminant removal is influenced by the presence of anions, NOM, and alcohols [41,145]. As a result, HVMs have attracted growing research interest and ongoing scientific debate over the past decade [138,146]. A key advantage of HVMs lies in the efficient generation of high-valence Fe species, which are produced from their corresponding low-valence forms (Fe(II) and Fe(III)) via a single-step, two-electron transfer process via Equation (22) [143]. Notably, compared to , PAA more effectively promotes the oxidation of low-valent iron to Fe(IV), due to its lower energy of the lowest unoccupied molecular orbital (LUMO) and weaker O–O bond strength, both of which facilitate enhanced electron transfer [39].
HVMs produced during PAA activation have been widely identified using a combination of probe compounds, such as dimethyl sulfoxide (DMSO) and methyl phenyl sulfoxide PMSO, with 18O-labeled (18O) [143,147], alongside various spectroscopic techniques. These identification methods have provided evidence that HVMs form in both homogeneous and heterogeneous PAA activation processes.
In 2019, Kim et al. [39] were the first to propose that Fe(IV) may play a pivotal role in the degradation of certain CECs, including NAP and BPA, within the Fe(II)/PAA system. This hypothesis was based on the substantial conversion of PMSO to PMSO2 and the minimal quenching effect observed in the presence of TBA at pH levels of 3, 5, and 7. Despite advances in the understanding of HVMs, the potential use of Co(IV) for wastewater remediation remained largely unexplored and unclear until 2021. In that year, Liu et al. [41] confirmed the formation of Co(IV) during PAA activation by Co(II) through PMSO-based probe tests, 18O isotope labeling experiments, and in situ Raman spectroscopy. This finding aligns with several studies that have highlighted the pivotal function of HVMs. For instance, in the /PAA/UVA process, the high-valent iron (IV)–oxo complex () was identified as the dominant RS. This species plays a critical role in the efficient and selective degradation of CECs. The research demonstrated that serves as the primary oxidant, driving the degradation process, while the generation of organic radicals and was found to be minimal in comparison [148]. Similarly, in the cobalt(II)/PMS system, the formation of plays a pivotal role in AOPs, further reinforcing the importance of HVMs as primary oxidants. It was observed that PMSO was readily oxidized to the corresponding sulfone (PMSO2) with a transformation ratio of ∼100% under acidic conditions, which strongly implied the generation of high-valent cobalt–oxo species instead of in the Co(II)/PMS process [149].
The interaction between HVMs and pollutants mainly occurs through three fundamental processes: hydrogen atom extraction [150], which involves the removal of a hydrogen atom from the pollutant molecule, frequently generating reactive intermediates; electron transfer [146], where the metal acts as an oxidizing agent by acquiring electrons from the pollutant; and oxygen atom transfer [151], whereby the metal aids in incorporating an oxygen atom into the pollutant, leading to its oxidation.
For the hydrogen atom extraction process involving aliphatic alcohol, Fe(IV) can remove the hydrogen atoms from the hydroxyl group, leading to the formation of an aldehyde or ketone, while Fe(II) is simultaneously formed through a two-electron transfer from Fe(IV). Additionally, the H atom on the α-C of RnCH–OH can be abstracted to form the radical intermediate R2C•–OH, which subsequently reacts with . The formed RC(OH)OO• intermediate undergoes an elimination reaction to produce R2C=O [150]. In the context of electron transfer, HVMs can react with compounds containing electron-donating groups via electrophilic addition pathways. For instance, Fe(IV) reacts with phenol by initially attacking the para-carbon of the aromatic ring. This electrophilic addition, followed by subsequent elimination, leads to the formation of hydroquinone [146].
Moreover, through an oxygen atom transfer mechanism, HVMs can oxidize sulfoxides into their corresponding sulfones. This pathway is particularly distinct, as sulfones are generally not produced in radical-driven reactions involving sulfoxides, highlighting the unique reactivity of HVMs in such transformations [152].
5. By-Products and Hazard Assessment of PAA-Based AOPs
Through the formation of RS, PAA-based AOPs have attracted increasing interest due to their efficiency in breaking down CECs. However, during oxidative reactions in water treatment, several by-products can be produced, some of which require assessment to ensure environmental safety. Among these, the primary by-products include carboxylic acids such as acetic acid [46]. The accumulation of acids in treated water can significantly lower pH, thereby affecting water quality and increasing the need for post-treatment neutralization.
Long-term acid accumulation can adversely affect aquatic ecosystems, impairing respiration and ion regulation in organisms, reducing biodiversity, and favoring acid-tolerant species. Chronic exposure may also result in sublethal effects, such as impaired reproduction in invertebrates, behavioral changes in fish, and reduced photosynthetic efficiency in algae. At the microbial level, while low concentrations of acetic acid may stimulate heterotrophic activity, elevated levels can inhibit sensitive populations and disrupt key processes such as nitrification. Moreover, due to its high biodegradability, acetic acid may increase biochemical oxygen demand (BOD), contributing to oxygen depletion in stagnant or eutrophic environments. When reused for irrigation, residual acetic acid may alter soil pH and hinder plant growth, particularly in alkaline or sensitive soils. Although acetic acid is not considered highly hazardous, it is regulated in effluent discharge standards to mitigate ecological acidification and oxygen stress.
In addition to acids, PAA disinfection also results in the formation of oxygenated by-products such as aldehydes and ketones, whose formation is directly related to the PAA dosage [153]. These compounds are biologically reactive and may persist under suboptimal degradation conditions. Formaldehyde, for example, is a probable human carcinogen (IARC Group 2A) and presents significant ecotoxicological risk [154]. It has an acute 96 h LC50 of 24 mg/L for Pimephales promelas and Oncorhynchus mykiss, and a LOAEL of 15 mg/kg/day in rodent studies, indicating both irritant and developmental effects [155]. Acetaldehyde, while of moderate toxicity, has an LC50 of 122 mg/L for Daphnia magna and has been associated with developmental toxicity under chronic exposure. Both aldehydes are capable of inducing oxidative stress and potential endocrine disruption and may bioaccumulate to a limited extent.
Importantly, aldehydes like formaldehyde may undergo further oxidation in the presence of excess PAA, forming carboxylic acids such as formic acid, introducing secondary environmental impacts [153].
Furthermore, in natural waters containing halide ions (e.g., Br− and Cl−), these can be oxidized by PAA into halogenated species such as hypohalous acids and halogen radicals [156]. These reactive intermediates can then react with organic matter to form halogenated by-products. Bromide ions are particularly concerning due to the significantly higher reaction rate constant with PAA compared to chloride [157]. Brominated organic by-products, such as bromoform, are persistent, bioaccumulative, and highly toxic. For instance, bromoform has an LC50 of approximately 5 mg/L for Daphnia magna and has been linked to reproductive and developmental toxicity in fish embryos.
To minimize the risks of secondary contamination, a comprehensive understanding of the formation mechanisms and influencing factors—such as PAA concentration, water matrix composition (especially halide content), activation method, and pH—is essential. Optimizing operational parameters and monitoring by-product profiles are crucial for ensuring the environmental safety of PAA-based AOPs over both short and long timescales.
6. Factors Affecting the Removal of CECs in PAA-Based Systems
6.1. Impact of Initial PAA Concentration
The initial dosage of PAA is a key factor affecting the removal efficiency of CECs in PAA-based processes. Generally, increasing the PAA concentration enhances the generation of RS in solution, thereby accelerating the degradation rate of CECs. Furthermore, higher PAA dosages increase the availability of oxidants. However, this benefit typically reaches a maximum at an optimal concentration, after which excessive PAA may actually hinder degradation performance. An excessive amount of PAA may act as a scavenger of reactive species. This phenomenon is often referred to as self-quenching. Kim et al. [39] observed that the degradation of micropollutants in the Fe(II)/PAA process at an initial pH of 3 increased from 69.7% to 100% over 60 min when the PAA concentration was raised from 50 μM to 500 μM, with [Fe(II)]0 fixed at 100 μM. However, further increasing the PAA concentration to 1000 μM reduced the degradation rate, thereby inhibiting micropollutant degradation. This decline can be attributed to the excessive PAA quenching of reactive radicals, which subsequently results in the production of less reactive species (Equations (23) and (24)) [39,57].
Kim et al. [20] also observed that in the Co(II)/PAA process, CBZ degradation significantly increased from 24.3% to 91.9% within 30 min as the PAA dosages were increased ([PAA]0 = 20, 50, 100, and 200 μM). This enhancement was reflected in the pseudo-first-order rate constant, which rose from (6.98 ± 0.04) × to (4.83 ± 0.67) × with increasing PAA concentrations. The observed trend can be attributed to the increased generation of RS in the solution at higher PAA concentrations. Zhang et al. [93] evaluated the effects of the initial PAA concentration on in the UV/PAA system for DCF degradation. They observed that increased gradually from 0.0967 to 0.182 min−1 as the PAA concentration was raised from 10 to 150 μM. However, increased proportionally with the rise in the PAA dose up to 50 μM, likely due to the enhanced production of and radicals. Unlike the linear increase, continued to rise at higher PAA concentrations, but the rate of increase slowed. This behavior is likely attributed to the dual effect of excess PAA scavenging radicals and competing with DCF for UV light absorption, thereby inhibiting direct photolysis.
When assessing the effects of PAA concentration on the removal of CECs in PAA-based systems, it is essential to maintain a constant pH value. Despite its importance, this detail is often overlooked in numerous research studies.
6.2. Effects of Catalyst Dosage
Degradation efficiency in AOPs can be more effectively enhanced by introducing a catalyst rather than simply increasing the oxidant concentration. Extensive research has demonstrated that higher catalyst dosages promote PAA activation, thereby improving the removal of CECs [39]. However, an excessively high catalyst concentration may lead to radical scavenging by the catalyst itself, ultimately reducing the degradation efficiency, which could either plateau or even decline. For example, Kim et al. [39] applied the Fe(II)/PAA process to the removal of CECs and investigated the effect of Fe(II) concentrations ranging from 20 to 500 μM. They observed a modest increase in the removal rate, from 63.1% to 75%, at concentrations between 20 and 100 μM. However, at 500 μM, the removal rate declined significantly to 46%. This finding suggests that excessively high Fe(II) concentrations can have a detrimental effect by depleting reactive radicals. A parallel trend was observed in the Co(II)/PAA process for CBZ degradation [20]. Increasing the concentration of Co(II) from 1 to 20 μM resulted in a linear increase in the degradation rate constant. However, when the Co(II) concentration exceeded 20 μM, the degradation rate was inhibited, likely due to radical scavenging by excess Co(II), which disrupted the degradation process’s efficiency.
On the other hand, for heterogeneous catalysts, increasing the catalyst dosage provides a greater number of available active catalytic sites, which significantly enhances the system’s ability to accelerate chemical reactions and improve overall efficiency. For example, Wang et al. [52] employed the /PAA system with 100 μM of PAA to degrade SMX. In the absence of the catalyst, the SMX concentration showed negligible variation. However, increasing the dose from 0.025 g·L−1 to 0.2 g·L−1 enhanced the SMX removal efficiency from 36.2% to 81.9% after 30 min, with increasing from 0.015 min–1 to 0.060 min–1. This improvement was attributed to the accelerated production of radicals via PAA decomposition at the active sites on the surface, while the adsorption of SMX on the surface remained minimal. An equivalent enhancement was also noted in the breakdown of Orange G (OG) by the /PAA process [68]. The degradation efficiency of OG increased from 0 to 79% as the dosage was increased from 0 to 0.01 g·L−1. Furthermore, even when the loading was increased from 0.01 to 0.2 g·L−1, the degradation continued to improve.
Both homogeneous and heterogeneous catalyst types can exhibit comparable behavior; although raising the catalyst concentration generally accelerates reactions, there is a threshold beyond which it may result in unwanted side effects. For instance, Wang et al. [62] observed that as the Fe(II)–zeolite dosage increased from 0.2 to 0.8 g·L−1, the SMX removal rate constant rose from 0.013 to 0.13 min−1. In contrast, when the catalyst dosage was further increased to 1.0 g·L−1, the rate constant decreased to 0.77 min−1. Similarly, Zhou et al. [65] utilized a cobalt ferrite combined with a PAA system to degrade the organic contaminant rhodamine B (RhB). Increasing the dosage from 0.3 to 0.5 g·L−1 enhanced the removal of RhB. However, when the dosage was further increased to 0.5–0.7 g·L−1, the degradation rate of RhB remained relatively stable. At a concentration of 1.0 g·L−1, the degradation rate was significantly inhibited, likely due to the agglomeration of magnetic nanoparticles.
6.3. Effects of pH
In PAA-based AOPs, pH plays a critical role, influencing the behavior of oxidants, catalysts, radical formation, and target contaminant interactions. A particularly important aspect is its impact on the acid–base equilibrium of PAA, which directly affects both its stability and reactivity. With a dissociation constant (pKa) of 8.2, the protonated form of PAA predominates at pH values below 8.2, whereas the deprotonated anionic form (PAA−) becomes dominant at higher pH levels.
Regarding the UV/PAA process, PAA− ( = 1.005 V) is a less potent oxidation agent than PAA [154]. Despite this, PAA− exhibits a greater molar absorption coefficient at 254 nm compared to PAA, with values of 41.6 M−1cm−1 and 8.0 M−1cm−1, respectively, indicating that the photolysis of PAA− and the generation of radicals occur more efficiently under alkaline conditions [155]. For instance, Zhang et al. [93] systematically studied the effects of pH on DCF degradation in the UV/PAA process and identified an optimal pH of 8.5. Within the investigated pH range of 3.0 to 8.5, the increased progressively from 0.096 to 0.1902 min−1, emphasizing the improved degradation efficiency under alkaline conditions. Nevertheless, operating under strongly alkaline conditions can compromise the effectiveness of the UV/PAA process, attributable to the decreased redox potential of the reactive , as described by the Nernst equation. This is further evidenced by Zhang et al. [93], who observed a reduction in the to 0.1536 min−1 at a pH of 11.0, compared to its value at a pH of 8.5.
Pertaining to the metal/PAA system, pH significantly influences the equilibrium states of both metal ions and PAA, thereby affecting their speciation and reactivity [156]. The interactions between transition metals and PAA depend on electron exchange, with the neutral form of PAA exhibiting a greater tendency to accept electrons compared to its deprotonated counterpart [157]. As mentioned earlier, PAA predominantly exists in its neutral form at acid to neutral pH levels. This aligns with studies on SMX degradation using the Co/PAA process, which demonstrate its pH dependence, with optimal performance observed at a near-neutral pH [77]. On the flip side, Liu et al. [41] demonstrated that under acidic conditions, the combination of Co(II) and PAA facilitates the generation of Co(IV) species, which are highly effective in degrading refractory organic contaminants, such as CBZ. In contrast, increasing the solution pH to neutral conditions, which are more commonly employed, appears to shift the mechanism, favoring the conversion of Co(IV) into oxygen-centered radicals.
Furthermore, throughout the process of heterogeneous PAA activation, the impact of pH varies depending on the type of catalyst employed, such as metal oxides. pH can significantly influence the surface charges of the metal oxide, which in turn affects their interaction with PAA. For instance, a neutral pH was found to be favorable for the degradation of OG in the /PAA system due to the favorable electrostatic interaction between PAA and the catalyst surface. Under acidic conditions, the excess ions form hydrogen bonds with the O–O fragment of PAA, hindering its interaction with the positively charged surface. Conversely, under alkaline conditions, electrostatic repulsion between the negatively charged oxide surface and PAA similarly reduces the catalyst’s effectiveness in interacting with PAA [68]. Similarly, metal–organic framework gels (MOGs) have emerged as highly effective catalysts for PAA activation, with performance is significantly influenced by pH. For example, Zheng et al. [158] investigated the FeMn bimetallic MOG with a hierarchical porous structure (FeMn13BTC)/PAA system for the degradation of ofloxacin (OFX). Their pH influence experiments revealed that the presence of ions hindered the system, indicating that OFX degradation was more effective under alkaline conditions. The isoelectric point of FeMn13BTC was determined to be at a pH of 2.81, meaning its surface becomes negatively charged at pH values above this threshold. However, OFX degradation under alkaline conditions was not suppressed by electrostatic repulsion between FeMn13BTC and PAA−. Instead, the enhanced degradation observed in alkaline conditions was attributed to increased PAA hydrolysis at higher pH levels, resulting in more efficient PAA activation based on the FeMn13BTC/PAA system. Concerning carbon-based catalysts, the application of PAA under activated biochar (ABC) catalysis exhibits distinct mechanisms. In this context, ABC, derived from nut shells, was employed as an eco-friendly catalyst for the degradation of sulfamethazine (SMZ). Zhang et al. [103] demonstrated that the effective degradation of SMZ at an acidic pH (<4.62) was driven by the electrostatic attraction between the positively charged ABC surface and the dissociated acidic form of SMZ (SMZ−). Furthermore, pH plays a crucial role in both the catalytic efficiency and adsorption behavior of carbon-based materials, thereby influencing their performance in the degradation process [103].
Understanding these pH-dependent behaviors is essential for choosing a suitable catalyst for environmental applications.
7. Challenges and Future Perspectives
Various strategies have been employed to enhance PAA activation for the efficient degradation of CECS. These processes can produce a diverse range of RS, including free radicals such as , , and , as well as non-radical species like 1 and HVMs. Each of these RS contributes to the degradation of CECs to varying extents, depending on the specific activation pathway and reaction conditions. However, despite the promising development potential of PAA-based AOPs in recent years, several critical challenges persist and must be thoroughly addressed in future research to fully realize their effectiveness and applicability.
- (1)
- A comprehensive study of the selective oxidation mechanisms of both non-radical and organic free radical pathways is essential. Non-radical mechanisms, such as 1 and HVMs, have emerged as key areas of interest. Crucial subjects for further investigation involve the precise detection and quantification of and , as well as exploring the oxidation capacity of organic radicals toward pollutants. Additionally, the differences in product accumulation and toxicity alterations between the free radical and non-free radical pathways for the same pollutant require further investigation.
- (2)
- Conducting parameter optimization techniques and their applications; creating analytical approaches for three-dimensional and other multidimensional factors is a key strategy for obtaining more sophisticated parameter optimization.
- (3)
- Investigating affordable and sustainable approaches for activating PAA, such as solar irradiation, holds significant potential for the degradation of CECs. The solar irradiation/PAA process offers advantages like easy accessibility, the absence of the need for additional chemicals, and its renewable nature, making it a promising avenue for further development and application.
- (4)
- Developing catalysts with enhanced catalytic efficiency: PAA-based AOPs primarily focus on metal-derived catalysts, with being the most commonly studied activator. While exhibits high catalytic performance at a neutral pH, its release into the environment poses a secondary source of pollution, necessitating careful consideration in future applications. Therefore, iron activation is considered a more environmentally friendly option. However, iron-based systems are highly pH-sensitive and involve additional chemicals for pH regulation. Moreover, although heterogeneous catalysts release metal ions at lower concentrations, these ions can still accumulate to levels exceeding water treatment standards. To mitigate these issues, encapsulating metal nanoparticles within a carbon or polymer layer offers a promising solution. This encapsulation enables the precise tuning of the electronic structure and work function, while leveraging the electron tunneling effect at the composite interface to optimize catalytic performance. Such advancements could enhance the efficiency and environmental compatibility of PAA-based AOPs, paving the way for more sustainable water treatment technologies.
- (5)
- Given that most current studies use ultrapure water matrices, and few explore real wastewater conditions, the influence of coexisting ions and natural organic matter remains poorly understood. Further research using real water matrices is critically needed to better predict process performance under practical conditions.
- (6)
- Additionally, while the laboratory-scale efficiency of PAA-based AOPs has been demonstrated, few studies have addressed the development of scalable reactor designs or process integration strategies for industrial applications. Future work should focus on conducting pilot-scale studies to evaluate the use of PAA-based AOPs in large-scale systems, focusing on managing multiple pollutants in real-world water treatment processes.
- (7)
- Additionally, efforts should be made to integrate PAA-based AOPs with technologies such as electrochemistry, membrane filtration, and biological treatments to improve overall efficiency.
Author Contributions
Conceptualization, B.B., S.C.E. and M.Á.S.; resources, M.Á.S.; writing—original draft preparation, B.B. and S.C.E.; writing—review and editing, S.C.E.; visualization, B.B. and M.Á.S.; supervision, M.Á.S.; project administration, M.Á.S.; funding acquisition, M.Á.S. All authors have read and agreed to the published version of the manuscript.
Funding
The authors acknowledge financial support from this research that has been financially supported by Xunta de Galicia (ED431C 2021-43).
Data Availability Statement
All data are contained within the article.
Conflicts of Interest
The authors declare the following financial interests/personal relationships which may be considered as potential competing interests: M.A. Sanromán reports financial support was provided by Xunta de Galicia.
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Figure 1.
Potential of PAA activation for achieving simultaneous pathogen inactivation and CECs degradation [19].
Figure 1.
Potential of PAA activation for achieving simultaneous pathogen inactivation and CECs degradation [19].
Figure 2.
Principal reactions involved in SMX removal within the Co/PAA system [77].
Figure 2.
Principal reactions involved in SMX removal within the Co/PAA system [77].
Figure 3.
Potential RS formed in PAA-based AOPs.
Figure 4.
Illustration of degradation spectra upon the addition of TBA to identify the presence of [21].
Figure 4.
Illustration of degradation spectra upon the addition of TBA to identify the presence of [21].
Table 1.
A summary of studies utilizing transition metals for PAA activation in the removal of CECs.
Table 1.
A summary of studies utilizing transition metals for PAA activation in the removal of CECs.
| CEC | Processes | [CEC]0 | Removal Efficiency (%) | PAA Dose | Catalyst Dosage | pH | Reaction Time | Ref. |
|---|---|---|---|---|---|---|---|---|
| P-arsanilic acid | Fe2+/PAA | 5 µM | 98% | 400 µM | 200 µM | 3 | 20 s | [63] |
| Bisphenol-A | Fe2+/PAA | 60 mg·L−1 | 100 | 526 µM | 400 µM | 3.5 | 10 min | [64] |
| Naproxen | Fe2+/PAA | 15 µM | 100 | 100 μM | 100 μM | 3–8.2 | 120 min | [39] |
| Rhodamine B | CoFe2O4/PAA | 20 mg·L−1 | 95 | 800 μM | 2131.1 μM | 7 | 10 min | [65] |
| Sulfamethoxazole | Fe2+-zeolite/PAA | 5 µM | 100 | 400 μM | 150 mg·L−1 | 7 | 50 min | [62] |
| Tetracycline | Fe0/PAA | 10 μM | 95 | 100 μM | 1074.4 µM | 3.5 | 30 min | [66] |
| Sulfamethoxazole | Co2+/PAA | 10 μM | 100 | 100 μM | 10 μM | 3.5 | 20 min | [41] |
| Carbamazepine | Co2+/PAA | 10 μM | 84 | 100 μM | 10 μM | 3.5 | 30 min | [41] |
| Bisphenol-A | Co2+/PAA | 10 μM | 100 | 100 μM | 10 μM | 3.5 | 15 min | [41] |
| Atrazine | Co2+/PAA | 10 μM | 20 | 100 μM | 10 μM | 3.5 | 30 min | [41] |
| Naproxen | Co2+/PAA | 15 µM | 100 | 100 μM | 10 μM | 3–8.1 | 3 min | [20] |
| Crystal Violet | Co2+/PAA | 0.06 mM | 67 | 200 μM | 10 μM | 7 | 60 min | [67] |
| Congo red | Co2+/PAA | 0.05 mM | 98 | 200 μM | 10 μM | 7 | 60 min | [67] |
| Acid orange 7 | Co2+/PAA | 0.05 mM | 92 | 200 μM | 10 μM | 7 | 60 min | [67] |
| Orange G | Co3O4/PAA | 0.05 mM | 100 | 6574.6 μM | 415.3 μM | 7 | 90 min | [68] |
| Sulfamethoxazole | CoFe2O4/PAA | 10 μM | 85 | 100–200 μM | 426.2 μM | 7 | 30 min | [52] |
| Sulfamethoxazole | LaCoO3/PAA | 50 μM | 100 | 263 μM | 660 μM | 7 | 60 min | [69] |
| Sulfamethoxazole | Co0/PAA | 5 μM | 99 | 50 μM | 1696.8 μM | 7 | 5 min | [30] |
| Sulfamethoxazole | Mn3O4/PAA | 1 μM | 100 | 1000 μM | 218.5 μM | 6.5 | 12 min | [70] |
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Bouzayani, B.; Elaoud, S.C.; Sanromán, M.Á. Current Progress in Advanced Oxidation Processes for the Removal of Contaminants of Emerging Concern Using Peracetic Acid as an Effective Oxidant. Catalysts 2025, 15, 469. https://doi.org/10.3390/catal15050469
AMA Style
Bouzayani B, Elaoud SC, Sanromán MÁ. Current Progress in Advanced Oxidation Processes for the Removal of Contaminants of Emerging Concern Using Peracetic Acid as an Effective Oxidant. Catalysts. 2025; 15(5):469. https://doi.org/10.3390/catal15050469
Chicago/Turabian StyleBouzayani, Bakhta, Sourour Chaâbane Elaoud, and Maria Ángeles Sanromán. 2025. "Current Progress in Advanced Oxidation Processes for the Removal of Contaminants of Emerging Concern Using Peracetic Acid as an Effective Oxidant" Catalysts 15, no. 5: 469. https://doi.org/10.3390/catal15050469
APA StyleBouzayani, B., Elaoud, S. C., & Sanromán, M. Á. (2025). Current Progress in Advanced Oxidation Processes for the Removal of Contaminants of Emerging Concern Using Peracetic Acid as an Effective Oxidant. Catalysts, 15(5), 469. https://doi.org/10.3390/catal15050469
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