New Insight into the Degradation of Sunscreen Agents in Water Treatment Using UV-Driven Advanced Oxidation Processes

Abstract

This study evaluates, for the first time, the effects of UV/PMS and UV/H₂O₂/PMS processes on the degradation of sunscreen agents in synthetic and natural water matrices and compares their effectiveness with the more conventional UV/H₂O₂. Investigations were conducted using a mixture of organic UV filters containing 4-methylbenzylidene camphor (4-MBC) and 2-ethylhexyl-4-methoxycinnamate. Among the investigated UV-driven AOPs, UV/PMS/H₂O₂ was the most effective in synthetic water, while in natural water, the highest degradation rate was observed during the degradation of EHMC by UV/PMS. The degradation of UV filters in the UV/PMS system was promoted by sulfate radical (68% of the degradation), with hydroxyl radical contributing approximately 32%, while both radical species contributed approximately equally to the degradation in the UV/H₂O₂/PMS system. The Vibrio fischeri assay showed an increase in inhibition (up to 70%) at specific stages of UV/H₂O₂ treatment when applied to natural water, which further decreased to 30%, along with an increase in UV fluence and progressive degradation. The Pseudomonas putida test recorded minor toxicity (<15%) after treatments. Magnetic biochar utilized in conjunction with UV-driven AOPs exhibited superior performance in eliminating residual contaminants, providing an efficient and sustainable approach to mitigate sunscreen agents in water treatment.

1. Introduction

Over the last few decades, climate change and ozone depletion events have consequently caused a higher level of human exposure to ultraviolet (UV) light. While short-term exposure to UV light might have positive effects on human health, such as facilitating the synthesis of vitamin D, prolonged or excessive exposure to UV radiation poses a significant health risk [1]. Solar UV radiation can cause various detrimental health impacts such as sunburns, photoaging, and long-term effects such as skin cancers [2,3]. In order to protect the skin from damage promoted by UV sun radiation, sunscreens and other personal care products (PCPs) containing a variety of inorganic and organic chemical ingredients, known as sunscreen agents or UV filters, have been developed and used since the late 19th century [4,5]. UV filters as sun-blocking agents are also frequently incorporated into different industrial products such as pesticides, detergents, food packaging, and rubbers, as well as paints and plastics, to prevent the photodegradation of polymers and pigments, thereby extending the shelf life of products [4,6].

Organic, or so-called chemical UV filters, are composed of aromatic structures that block solar UV light via the absorption of UV-A (320–400 nm) or shorter wavelength UV-B (280–320 nm) radiation, while inorganic UV filters not only absorb UV light but also reflect and scatter UV radiation due to physical phenomena [5,7,8]. Organic UV filters can be classified as para-amino benzoates, salicylates, cinnamates, benzophenones, dibenzoyl methanes, and camphor derivates [9]. To achieve the desired level of protection against UV regions of solar radiation, many sunscreens and PCPs employ a combination of different families of UV filters [9]. Among them, 4-methylbenzylidene camphor (4-MBC) and 2-ethylhexyl-4-methoxycinnamate (EHMC) are common UV filters in sunscreens and PCPs.

UV filters enter the aquatic environment through direct routes such as recreational activities (e.g., washed from the swimmers) or indirectly via wastewater treatment plant effluents [10,11]. Due to their extensive use, UV filters are found ubiquitously and detected in various water matrices including freshwater, marine environments, swimming pool water, and even drinking water [6,12]. Surface water concentrations of 4-MBC range from 5 to 400 ng/L, and those of EHMC from 10 to 800 ng/L, varying seasonally with higher levels observed in summer [4,6]. Groundwater contamination occurs through the infiltration of surface water and landfill leachate, with detected levels of 4-MBC and EHMC reaching up to 100–200 ng/L [12], while wastewater treatment plants discharge up to 2000–5000 ng/L 4-MBC and EHMC, respectively [10,11]. PCPs containing sunscreen agents pose a potential risk to the environment and aquatic biota, thereby potentially impacting human health. This class of contaminants has been linked to numerous adverse toxicological effects, including the potential to disrupt the endocrine system and impact reproduction and development. Consequently, they are considered environmental contaminants of emerging concern (CECs). Most environmental research has been focused on the potential impact of UV filters on marine environments, while research dealing with their impact on freshwater and human health is still scarce [4,6]. Certain UV filters are listed on the watchlist of substances to be monitored in EU surface waters, supporting future prioritization.

Conventional technologies are not designed to fully eliminate CECs, including UV filters, making their removal very challenging. To enhance the protection of aquatic ecosystems and drinking water sources, advanced water treatment, including advanced oxidation processes (AOPs), should overcome the gaps in knowledge and eliminate challenging contaminants in water/wastewater treatment [13,14,15]. The realm of AOPs encompasses various methods, including ozonation, ozonation combined with hydrogen peroxide (H₂O₂), Fenton and analogous reactions, TiO₂-based photocatalysts, sonolysis, electrochemical oxidation, and assorted amalgamations thereof [16,17,18]. The utilization of UV irradiation alongside various catalysts has gained increasing interest, enabling the activation of oxidizing agents like H₂O₂, persulfate (PS, S₂O₈²⁻)/peroxymonosulfate (PMS, HSO₅⁻), peracetic acid (PAA, CH₃CO₃H), chlorine, and ozone (O₃), which results in the production of highly reactive oxygen species (ROS) [19]. H₂O₂ and PMS are frequently employed as sources for generating hydroxyl (HO^•) and sulfate (SO₄^•−) radicals. The radicals HO^• and SO₄^•−, characterized by redox potentials (E°) of 1.9–2.85V and 2.5–3.1 V, respectively, exhibit robust oxidizing properties [20,21]. They can efficiently oxidize and break down a wide range of contaminants in water and wastewater, with a high reaction rate and without selectivity. Furthermore, these oxidizing agents possess environmentally friendly characteristics and have extensive application in green engineering practices [22,23,24]. Recently, compounds like periodate have been utilized in AOPs combining with UV, visible light, and/or catalytic systems to produce various oxidizing substances, including IO₃^•, IO₄^•, HO^•, IO₃⁻, O(³P), H₂O₂, and O₃ [25,26,27,28]. The efficiency of AOPs largely depends on parameters such as reaction time, UV fluence, initial concentrations of the oxidant and target pollutant, pH, temperature, and the presence of inorganic anions and natural organic matter, requiring process optimization for site-specific application [29].

In water treatment, AOPs are frequently combined with adsorption processes. Compared with conventional activated carbon application, biochar-based adsorbents, derived from biomass pyrolysis, offer a novel and sustainable alternative [30,31]. Biochar production aligns with principles of environmental sustainability, utilizing renewable biomass sources and potentially contributing to carbon sequestration. However, the low particle size of biochar, along with its lower density, poses challenges in terms of regeneration, separation, and recovery following adsorption, which could impede its recycling potential and industrial usage [32,33]. To address these challenges, researchers have begun synthesizing magnetically recyclable biochar, known as magnetic biochar, by incorporating iron oxide (e.g., Fe₃O₄ and Fe₂O₃). Magnetic biochar has demonstrated good adsorption performance for removing various organic pollutants such as antibiotics, organic dyes, pesticides, and organochlorine compounds [34,35,36].

In recent years, the removal of organic UV filters containing 4-MBC and EHMC has been investigated using various methods, including ozonation, UV and solar irradiation, the Fenton and photo-Fenton processes, the UV/H₂O₂ process, the UV/persulfate process, and the UV/TiO₂ process, as summarized in Table 1 [9,37,38,39,40]. However, to date, there have been no published data regarding the degradation of 4-MBC and EHMC in a mixture using the novel UV/PMS process. Additionally, the effects of synergism between oxidants H₂O₂ and PMS combined with the UV-C light on the degradation of sunscreen agents in water remain underinvestigated. To bridge this gap, this study deals with the degradation of a mixture of camphor- and cinnamate-derivative UV filters for the first time using the novel UV/PMS process and UV-AOPs employing combined oxidants (UV/PMS/H₂O₂). The degradation efficacy of the utilized sulfate-radical-based AOPs was compared with the more conventional UV/H₂O₂ process. Special emphasis was placed on elucidating water matrix effects by detecting the dominant radical species involved in the degradation mechanism. Adsorption treatment using newly synthesized magnetic biochar was applied for the first time to mitigate residual organic UV filters and possible transformation products formed during the AOPs in water treatment. Furthermore, water quality was assessed using toxicity testing of water treated with UV-driven AOPs.

2. Materials and Methods

2.1. Chemicals

The main chemical reagents employed in this experiment were 4-methylbenzylidene camphor (4-MBC) as a certified reference material, TraceCERT^® (C₁₈H₂₂O, CAS No. 36861-47-9, Sigma Aldrich, St. Louis, MO, USA), 2-ethylhexyl-4-methoxycinnamate (EHMC) of analytic standard (C₁₈H₂₆O₃, CAS No. 5466-77-3, Sigma Aldrich, St. Louis, MO, USA), hydrogen peroxide (H₂O₂, 30% w/w, POCH, S.A., Gliwice, Poland), potassium peroxymonosulfate (2KHSO₅·KHSO₄·K₂SO₄, CAS No. 70693-62-8 available as Oxone, Sigma Aldrich, St. Louis, MO, USA), tert-butyl alcohol ACS reagent (C₄H₁₀O, CAS No. 75-65-0, Sigma Aldrich, St. Louis, MO, USA), sodium thiosulfate pentahydrate as ReagentPlus^® (Na₂S₂O₃·5H₂O, CAS No. 10102-17-7, Sigma Aldrich, St. Louis, MO, USA), and sodium chloride ACS reagent (NaCl, CAS No. 7647-14-5, Sigma Aldrich, St. Louis, MO, USA). Iron(II) sulfate heptahydrate (FeSO₄·7H₂O, CAS No. 7782-63-0, POCH S.A., Gliwice, Poland) and iron(III) chloride hexahydrate (FeCl₃·6H₂O, CAS No. 10025-77-1, POCH S.A., Gliwice, Poland) were used for the preparation of magnetic biochar. Methanol (CH₃OH, CAS No. 67-561), hexane (C₆H₁₄, CAS No. 110-54-3), and acetonitrile (CH₃CN, CAS No. 75-05-8) were purchased from J.T. Baker^® (Phillipsburg, NJ, USA). Ultrapure water obtained from a LABCONCO (Water Pro RO/PS Station, Kansas City, MO, USA) system (water of American Society for Testing and Materials type I quality, dissolved organic carbon < 0.5 mg/L; electrical conductivity 0.055 µS/cm) was used for the preparation of all synthetic water matrix and working solutions. Stock solution of each UV filter (1 mg/mL) was prepared in acetonitrile.

2.2. Water Samples

Synthetic water matrix (ultrapure water without the addition of interfering agents) and natural water were spiked with a mixture of 4-MBC and EHMC in order to obtain the final concentration of about 100 µg/L in water. The characteristics of the natural water sample used for the investigation included pH values of 7.81 ± 0.23, an electrical conductivity of 358 ± 22 μS/cm, a UV₂₅₄ of 0.075 ± 0.01 cm⁻¹, a UV₂₇₈ of 0.075 ± 0.01 cm⁻¹, an SUVA of 2.69 ± 0.13 L m⁻¹ mg⁻¹, an ammonia content of 0.172 ± 0.03 mg N/L, a nitrate content of 0.775 ± 0.10 mg N/L a nitrite content of 0.015 ± 0.02 mg N/L, an orthophosphate content of 0.012 ± 0.01 mg P/L, a bicarbonate content of 100.6 ± 0.25 mg HCO₃⁻/L, a hardness of 192 ± 14 mg CaCO₃/L, a chloride content of 14.8 ± 0.3 mg/L, and a sulfate content of 36.5 ± 2.70 mg/L.

2.3. UV-Driven AOP Water Treatment

The degradation of 4-MBC and EHMC were utilized in a photochemical reactor with a volume of 700 mL and a low-pressure lamp (16 W TUV Philips., UV-technik Speziallampen GmbH, Ilmenau, Germany) emitting monochromatic radiation at 253.7 nm. Via the application of chemical actinometry with potassium ferrioxalate, the photon flux (2.33 × 10⁻⁶ Einsteins s⁻¹) and incident light flux (1.5 mW cm⁻²) were determined. In our previous investigation by Đurkić et al. [41], a detailed description of the procedure and schematic of the photoreactor was provided. The influence of different oxidant concentrations ([H₂O₂] = [HSO₅⁻] = 0.03 and 0.3 mM) and UV fluence values (33–1400 mJ/cm²) on sunscreen agent degradation was investigated. The residues of oxidant agents were quenched with 0.1 M Na₂S₂O₃. Each set of experiments was performed in triplicate and the mean values were used for data interpretation and presentation.

2.4. Preparation of Magnetic Biochar

The modification of biochar with magnetic nanoparticles began by suspending 0.5 g of pristine biochar (obtained from corn biomass) in water and subjecting it to ultrasonication to prevent agglomeration. Subsequently, 3 g of FeCl₃·6H₂O and 1.4 g of FeSO₄·7H₂O were added to the biochar suspension. Then, 5 M NaOH was introduced dropwise into the mixture until the pH reached 9–10, resulting in the formation of a black precipitate. This black precipitate, representing magnetic biochar, was then separated from the water using an external magnet, washed several times with deionized water until the supernatant reached a neutral pH, and finally dried at ambient temperature. The resultant material was further characterized and investigated as a potential adsorbent for tertiary/quaternary treatment.

2.5. Toxicity Evaluation

Standardized assays with Pseudomonas putida and Vibrio fischeri were used to assess the toxicity of the target compounds and their possible degradation products in water treated by AOPs. In the Pseudomonas putida growth inhibition test, the inhibitory effect of water samples was determined by comparing cell growth with varying dilutions of test samples to the cell growth of a culture obtained under the same conditions but without test samples. The percentage of growth inhibition (I%) was calculated based on growth in different sample concentrations compared to the control without the sample (ISO 10712:1995) [42]. The toxicity of water samples towards the luminescence inhibition rate of Vibrio fischeri culture was measured using the LUMIStox 300 Bench Top Luminometer with LUMIStherm thermostat and LUMISsoft4 PC software (Hach-Lange GmbH, Düsseldorf, Germany,) based on the standard method ISO 11348-1:2007 [43]. The luminescence intensity of Vibrio fischeri was determined by mixing varying concentrations of the sample with reconstituted bacterial suspension and sodium chloride solution, as prescribed by the standard. After exposing Vibrio fischeri to the untreated and treated samples for 30 min, the inhibition of luminescence was calculated as the percentage of inhibition effect (H%) based on luminescence values in different sample concentrations and control.

2.6. Instrumental Analysis

After UV-AOPs treatments, a mixture of sunscreen agents was extracted from the reaction solutions using a liquid–liquid extraction with hexane prior to the chromatographic analysis. For the detection of the 4-MBC and EHMC, a gas chromatography with mass spectrometry (Agilent Technologies 7890A Gas chromatograph/5975C Mass spectrometer (GS-MS), Santa Clara, CA, USA) with a DB-5MS capillary column ((30 m × 25 μm × 0.25 µm) J&W Scientific, Santa Clara, CA, USA)) was used. The chromatography conditions were optimized for this study based on a procedure outlined in the literature [44,45]. The chromatography conditions were as follows: the oven was set initially at 70 °C (0 min), then ramped up to 250 °C at a rate of 30 °C/min (for 5 min), and finally to 280 °C at a rate of 30 °C/min (0 min). The ion source temperature was maintained at 230 °C, while the quadrupole temperature was maintained at 150 °C throughout the analysis in electron impact mode (70 eV). The transfer line temperature was set at 280 °C, and a solvent delay of 4.50 min was implemented. Splitless injection was employed with an injection volume of 2 µL. The analysis was conducted in SIM/SCAN mode. The quantitation ion (m/z) for 4-MBC was 254 with confirmation ions (m/z) 171 and 128, while for EHMC, the quantitation ion (m/z) was 290 with confirmation ions (m/z) 178 and 161. The recovery was in the range of 86% to 114% with relative standard deviations (RSDs) <15%. The method detection limit (MDL) was determined to be 50 ng/L.

The pH value of the water was measured using a pH/ION 735 instrument (Model 04520006; WTW, Weilheim, Germany). Water electrical conductivity was measured using an electroconductometer (Hanna Instruments, Model HI 933000; TX, Cluj-Napoca, Romania). Total organic matter content in natural water samples was determined by measuring total organic carbon (TOC) and dissolved organic carbon (DOC) using a TOC analyzer ElementarLiqui TOC II (Model 35072028; Elementar, Langenselbold, Germany). UV absorbance at 254 nm was determined by measurement using a CINTRA 1010, GBC Scientific Equipement spectrophotometer. Specific UV absorbance (SUVA) was calculated based on UV₂₅₄ and DOC values, according to the standard method [46]. The concentration of anions in natural water was determined using the ion chromatography system Dionex IC 3000 (Santa Clara, CA, USA).

The multi-point BET-specific surface areas, mesopore volumes, and micropore volumes of both the pristine biochar and the magnetic biochar were determined using the AutosorbiQ Surface Area Analyzer (Quantachrome Instruments, Boynton Beach, FL, USA). The specific surface areas were computed by utilizing the multi-point Brunauer–Emmett–Teller (BET) method, while the determination of meso and micro pore volumes was accomplished using the desorption Barrett–Joyner–Halenda (BJH) isotherms and the t-test method, respectively. X-ray diffraction (XRD) patterns were collected using a Philips PW automated X-ray powder diffractometer located in the USA. Fourier transform infrared (FTIR) spectra of the magnetic nanoparticles were obtained using an infrared spectrophotometer: the FTIR Nexus 670, manufactured by Thermonicolet in the USA. The point of zero charge (pHpzc) of the magnetic nanoparticles was determined by following the method outlined by Šolić et al. [47].

3. Results and Discussion

3.1. Degradation of Sunscreen Agents in Ultrapure Water by UV-Driven AOPs

To better elucidate the behavior of UV filters during the different AOPs, the physico-chemical characteristics of the target contaminants subjected to the water treatment are provided in Table 2 and used for further discussion. Figure 1 illustrates the degradation of sunscreen agents using various processes, including the application of H₂O₂, PMS, and UV-C photolysis alone as control processes, and a combination of UV irradiation with oxidant agents. Additionally, the figure shows the results of the investigation of the synergistic effect of combined H₂O₂ and PMS activated under the UV-C light in order to evaluate whether the interaction between the two oxidants can enhance the degradation of target contaminants.

It is important to note that both the (E) and (Z) geometric isomers of the olefinic UV filters need to be considered during water treatment. These compounds demonstrate rapid photoisomerization between their (E) and (Z) forms, eventually reaching a pseudo-equilibrium state [9]. Thus, the results obtained are presented as the sum of both geometric isomers. The obtained results indicate that neither of contaminants were degraded (less than 5%) when treated solely with H₂O₂ and PMS (0.03 mM) or with a combination of oxidants ([PMS]₀ = [H₂O₂]₀ = 0.015 mM). However, applying UV-C photolysis achieved significantly higher efficiency in degrading UV filters (about 53% at a UV fluence of 1400 mJ/cm²) compared to H₂O₂ and PMS alone or a combination of oxidants (PMS/H₂O₂). Moreover, the combined action of UV irradiation and oxidant notably enhanced UV filter degradation due to the activation of H₂O₂ and PMS and the generation of HO^• and/or SO₄^•− through chain reactions that interact with organic contaminants (reactions 1–3) [50,51].

H₂O₂ + hv → 2HO^•(1)

HSO₅⁻ + hv → SO₄^•− + HO^•(2)

SO₄^•− + H₂O → HO^• + SO₄²⁻(3)

HSO₅⁻ + H₂O₂ → SO₄^•− + 3HO^•(4)

As shown in reaction (4), PMS and H₂O₂ can be activated not just through excitation by UV irradiation (reactions 1–3) but also through their mutual reaction, leading to the formation of SO₄^•− and HO^• radicals.

The reaction kinetics of PMS activation by H₂O₂ (k = 1.7 × 10⁷ M⁻¹s⁻¹) are slow [52,53]; however, combining them under UV-C light significantly enhances sunscreen agent degradation compared to their isolated effect or UV-C-mediated activation.

The degradation efficacy of sunscreen agents in ultrapure water decreased according to the following trend: UV/PMS/H₂O₂ ˃ UV/H₂O₂ ˃ UV/PMS, requiring significantly lower UV fluence (200 mJ/cm²) compared to the UV photolysis alone. The obtained results indicated that the simultaneous presence of SO₄^•− and HO^• in the reaction solution was beneficial for improving the degradation efficiency of 4-MBC and EHMC. However, only a few studies have explored this system’s application to wastewater treatment [53,54,55], necessitating additional research to thoroughly understand the reactions involved and assess its feasibility [52].

The pseudo-first-order kinetic model was employed to describe the degradation kinetics of sunscreen agents during UV-based AOPs (Figure 2). The rate constants (k) were determined based on fluence, calculated as the slope of the natural logarithm of the ratio of initial to final concentrations of contaminants ([C]/[C₀]) against UV fluence (mJ/cm²), where [C₀] and [C] represent the initial and final concentrations of sunscreen agents in the water samples [56]. The degradation rate constant for sunscreen agents in ultrapure water using UV-C photolysis alone was 0.501 ± 0.032–0.528 ± 0.028 × 10⁻³ cm² mJ⁻¹ (R² = 0.997). During the UV/H₂O₂, UV/PMS, and UV/PMS/H₂O₂ processes, the k values significantly increased, ranging from 8.54 ± 0.6 to 17.8 ± 0.22 × 10⁻³ cm² mJ⁻¹ (R² = 0.918–0.993).

The obtained results indicate that the highest k values were observed for EHMC degradation by the UV/PMS/H₂O₂ system. The co-addition of PMS and H₂O₂ to the reaction medium under UV-C light accelerated the contaminants reaction rate by 1.5 times compared to the UV/H₂O₂ process and/or the UV/PMS process. The effectiveness of contaminant removal in an AOP depends on the type and quantity of radicals, the reactivity of generated radicals with the contaminant, and the nature of the degradation intermediates [53].

Considering that the investigation of the degradation of the contaminants was conducted in a mixture, the potential influence of competition reactions should not be overlooked, as this may arise during such an examination. Theoretically, more reactive pollutants may scavenge ROS, reducing their availability for the oxidation of other pollutants. Additionally, the degradation rates of organic compounds in UV-driven AOPs can vary depending on their molecular structure. In the chemical structure of EHMC, a 2-ethylhexyl group is present, which refers to a branched alkyl group, along with a 4-methoxycinnamate group (cinnamate group or phenylacrylic acid), with a methoxy group (-OCH₃) substituted at the para position of the aromatic ring [57]. On the other hand, 4-MBC has a benzylidene group and camphor structure, which consists of a bicyclic ketone comprising a fused cyclohexane and cyclopentane ring system. EHMC degraded slightly faster, presumably due to its ester functional group (C=O-O-R), which was more susceptible to photodegradation under UV-C irradiation compared to the ketone functional group of 4-MBC. The ester group can undergo photolysis, leading to the cleavage of the carbon–oxygen bond and subsequent degradation of the compound. Also, the reactivity of organic compounds can be influenced by the presence of various substituents in their molecular structure. In the case of EHMC, the ethylhexyl group and the methoxy group may increase the susceptibility of the molecule to photodegradation through pathways such as hydrogen abstraction or electron transfer reaction. 4-MBC contains aromatic rings in its structure, which can provide some degree of stability against photodegradation. Aromatic compounds often exhibit resonance stabilization, which can make them less reactive toward UV-induced degradation.

The rate of free radical generation through UV/peroxide processes primarily depends on the molar extinction coefficients (ε) and photolysis quantum yields (Φ) of peroxides. The data revealed that the ε values for H₂O₂ and its dissociated form HO₂⁻ were 19.6 and 229 M⁻¹cm⁻¹ were higher than those for HSO₅⁻ (the monovalent form of PMS) and SO₅²⁻, the divalent form of PMS (13.8 and 149.5 M⁻¹cm⁻¹, respectively). Furthermore, the quantum yield of H₂O₂ photolysis at 254 nm was determined to be Φ = 0.5, whereas the quantum yield of PMS photolysis at the same wavelength was measured as Φ = 0.52 [58]. Consequently, the slightly better performance of the UV/PMS/H₂O₂ system can be attributed to the efficient absorption of UV light facilitating H₂O₂ and PMS photolysis, which leads to an increased amount of ROS. It has also been noted that H₂O₂ tends to undergo homolytic cleavage more readily than PMS, leading to the assumption that more hydroxyl radicals were generated in the combined system, which was responsible for the degradation of target contaminants [59,60].

3.2. Identification of Radical Species Involved in Sunscreen Agent Degradation by AOPs

To further explore the roles of HO^• and SO₄^•− in the degradation of sunscreen agents, target pollutants were degraded by UV-driven AOPs in the presence of radical scavengers. Methanol (MeOH) and tert-butyl alcohol (TBA) were employed as scavengers for HO^• and SO₄^•− due to their high reactivity towards these species. Methanol exhibits a rapid reaction rate with both radicals. TBA reacts rapidly with HO^• radical, which is three orders of magnitude faster than with SO₄^•− (reactions 5–8). Therefore, TBA was commonly employed as a hydroxyl scavenger, whereas MeOH was utilized to efficiently quench these two radicals by adjusting the appropriate dosage [57,61,62,63].

(5)${{\mathrm{SO}}_4}^{•-}+\mathrm{MeOH} \to products (k_{\raisebox{1ex}{$SO{4}^{•-}$}\!\left/ \!\raisebox{-1ex}{$MeOH$}\right.}=1.1\times {10}^7 {\mathrm{M}}^{-1}{\mathrm{s}}^{-1})$

(6)${\mathrm{HO}}^{•}+\mathrm{MeOH} \to products (k_{\raisebox{1ex}{$H{O}^{•}$}\!\left/ \!\raisebox{-1ex}{$MeOH$}\right.}=9.7\times {10}^8 {\mathrm{M}}^{-1}{\mathrm{s}}^{-1})$

(7)${{\mathrm{SO}}_4}^{•-}+\mathrm{TBA} \to products (k_{\raisebox{1ex}{$SO{4}^{•-}$}\!\left/ \!\raisebox{-1ex}{$TBA$}\right.}=\left(4.0–9.1\right)\times {10}^5 {\mathrm{M}}^{-1}{\mathrm{s}}^{-1})$

(8)${\mathrm{HO}}^{•}+\mathrm{TBA} \to products (k_{\raisebox{1ex}{$H{O}^{•}$}\!\left/ \!\raisebox{-1ex}{$TBA$}\right.}=\left(3.8–7.6\right)\times {10}^8 {\mathrm{M}}^{-1}{\mathrm{s}}^{-1})$

Figure 3 and Figure 4 illustrate the quenching effect of MeOH and TBA on the degradation of target contaminants by the UV/H₂O₂, UV/PMS, and UV/PMS/H₂O₂ systems compared to the control treatments without MeOH and TBA addition. Other reactive species could also be generated in UV-driven AOPs, including singlet oxygen (¹O₂) and superoxide anion (O₂⁻), but cannot be quenched by the application of MeOH and TBA [61,62,63]. Although ¹O₂ and O₂⁻can also participate in oxidation processes, they do so with a significantly lower oxidation potential and reactivity compared to HO^• and SO₄^•− radicals. Hence, in this study, the emphasis was placed on identifying the main radical species of HO^• and SO₄^•− in UV-driven AOPs.

The degradation of both sunscreen agents in the UV/H₂O₂ process significantly decreased from 91–93% to 14–22% after adding 50 mM of MeOH and TBA (Figure 3a and Figure 4a). The obtained experimental results indicate that both quenchers were effective HO^• scavengers in the UV/H₂O₂ process, suggesting that generated HO^• radicals were almost completely quenched using 50 mM of MeOH and TBA (k value significantly decreased by approximately 90%). Based on these results, HO^• radicals play a key role in the degradation of UV filters in the UV/H₂O₂ system. During the UV/PMS process, the addition of 50 mM MeOH decreased the degradation efficiency of 4-MBC and EHMC from 86 to 72% and 44%, respectively. In the presence of 50 mM TBA, the degradation of both compounds decreased to 40–44% (Figure 3b and Figure 4b). This notable distinction and the impact of quenchers imply the involvement of both SO₄^•− and HO^• radicals in the degradation of sunscreen agents by the UV/PMS process. Furthermore, the lower inhibition degree observed with the addition of MeOH suggests the predominant involvement of SO₄^•− radicals in the degradation of 4-MBC by the UV/PMS process. Also, a series of tests were conducted to identify the predominant radicals in the UV/PMS/H₂O₂ system (Figure 3c and Figure 4c). The obtained results indicate that the addition 50 mM MeOH/TBA does not significantly inhibit the process, providing similar degradation efficacy as achieved during the UV/H₂O₂ and UV/PMS processes. Increasing the amount of TBA (200 mM) in UV/PMS/H2o2 resulted in a decrease in the degradation of target contaminants from 98% to 66%. However, under the same conditions with the addition of MeOH, the degree of degradation decreased to 56% compared to the control process. Accordingly, HO^• was slightly more significant in UV/PMS/H₂O₂ system, although SO₄^•− simultaneously contributes to the degradation of target contaminants. To date, there are limited literature data discussing the simultaneous use of PMS and H₂O₂ in combination with UV radiation for water treatment or focusing on the identification of dominant radicals in such systems. Zhong et al. [54] found a predominant presence of HO^• in their study on the removal of chloramphenicol. However, it is important to note that their study was based on an examination of an H₂O₂/PMS double-oxidation system catalyzed by pipe deposits, so the results may not be directly comparable to our study. Amanollahi et al. [53] observed a higher significance of HO^• radicals during the removal of ammonia nitrogen using a VUV/H₂O₂/PMS system, although they also reported a notable contribution from SO₄^•−.

The relative contributions of SO₄^•− and HO^• radicals in the sulfate-based AOPs can be calculated using Equations (9) and (10) [64].

(9)${HO}^{•} contribution (\%)=\frac{k with H{O}^{•}scavenger}{k without scavenger}\times 100$

SO₄^•−contribution (%) = 100 − HO^• contribution (%)(10)

where [k with HO^• scavenger] is the obtained pseudo-first-order degradation rate constant with the scavenger added and [k with without scavenger] is the pseudo-first-order degradation rate constant in the UV/PMS system. From the quenching experiments, the contributions to the degradation of 4-MBC in the UV/PMS were as follows: SO₄^•− accounted for 68.5%, and HO^• for about 31.5%. Hsieh et al. [40] investigated the dominant radical species in the degradation of 4-MBC when using the UV/persulfate process. The quenching experiment was performed by adding 1.26 mM TBA into the solution, suggesting that HO^• accounted for approximately 30%, while SO₄^•− accounted for most of the degradation of 4-MBC (~70%) as the major radical species in the UV/persulfate process.

3.3. Degradation of Sunscreen Agents in Natural Water by UV-Driven AOPs

Natural waters are complex mixtures that typically include natural organic matters (NOM) and inorganic ions such as HCO₃⁻/CO₃²⁻, NO₃⁻, Cl⁻, etc., which can interact with radicals, thus influencing the degradation of organic compounds [65,66]. The influence of water matrix constituents on the oxidation of sunscreen agents remains poorly elucidated. Traditionally, it has been assumed that these constituents adversely affect the reaction system. Nevertheless, a few studies have indicated that the presence of water matrix components can enhance the efficacy of sulfate-radical-based AOPs (SR-AOPs) while reducing the efficiency of AOPs relying on hydroxyl radicals under real water treatment conditions. Moreover, the composition of humic acid as a representative NOM varies significantly depending on its source, which can introduce complexities in its impact on the oxidation of the target compound. This influence may vary based on the source of NOM, the oxidation method employed for the target compound, and the presence and composition of various other water matrix components [19,64].

The investigation of the effect of UV-driven AOPs on the degradation of sunscreen agents in natural water is shown in Figure 5. By applying the UV photolysis alone, the degradation efficiency of target contaminants was up to approximately 45% at a UV fluence of 1400 mJ/cm². Seo et al. [9] investigated the degradation of eight organic UV filters, including 4-MBC and EHMC, using a direct UV_254nm photolysis in drinking water and wastewater, where a similar the degree of degradation was obtained (about 38% at a UV fluence of 1500 mJ/cm² and pH 7).

The addition of oxidants H₂O₂ and PMS significantly improves the degradation of sunscreen agents in natural water compared to UV photolysis alone, which can be attributed to the photolytic decomposition of H₂O₂ and PMS. The degree of degradation of sunscreen agents in natural water at the initial oxidant concentration of 0.03 mM and with a UV fluence of 200–1400 mJ/cm² during UV/H₂O₂ and UV/PMS ranged from 36 to 79% and from 45 to 92%, respectively. Similar results were obtained when PMS and H₂O₂ were simultaneously added at a concentration of 0.015 mM in the combined UV/PMS/H₂O₂ system. The presence of both PMS and H₂O₂ in the reaction medium led to their simultaneous activation by UV light and the generation of both SO₄^•− and HO^• radicals. However, the results indicated that in the UV/PMS/H₂O₂ process, the efficiency of degradation (34–74%) was almost identical to that during the UV/H₂O₂ treatment. An increase in the initial concentration of the applied oxidants (0.3 mM) resulted in a further increase in the degradation efficiency, amounting to 43–83% in the UV/H₂O₂ system and 53–97% in the UV/PMS system. In the UV/PMS/H₂O₂ system, increasing the concentration of oxidants to 0.15 mM resulted in the same degradation trend as observed in the UV/H₂O₂ process. Therefore, to evaluate the suitability of the combined UV/PMS/H₂O₂ process for the treatment of natural water and to demonstrate the effectiveness of dual oxidation activated by the UV light, further research should focus on examining the concentration of oxidants and their molar ratio in the system.

The presented results indicate that EHMC degrades somewhat faster than 4-MBC, whereby a removal efficiency of EHMC of over 90% can be achieved by applying the UV/PMS process (0.3 mM PMS; 1000 mJ/cm²). Additionally, the efficiency of target contaminant degradation was notably hindered in natural water compared to ultrapure water, which can be attributed to interference from organic matter and inorganic ions. Nevertheless, the UV/PMS process enables a high efficiency of EHMC degradation, suggesting its effectiveness and moderate resilience against the water matrix components’ influence. Seo et al. [9] indicate that the application of UV/H₂O₂ (10 mg H₂O₂/L and the UV fluence of 100–1500 mJ/cm²) reduces 4-MBC and EHMC by approximately ≥ 90% in drinking water and wastewater treatment.

The degradation rate constants of UV filters in natural water during the UV photolysis were in the range of 0.417 ± 0.05–0.424 ± 0.02 × 10⁻³ cm² mJ⁻¹ (R² = 0.986) (Figure 6). The obtained k values for the UV/H₂O₂, UV/PMS, and UV/PMS/H₂O₂ processes were significantly higher and ranged from 0.815 ± 0.08 to 2.36 ± 0.08 × 10⁻³ cm² mJ⁻¹ (R² = 0.906–0.969).

The highest degradation rate was observed for EHMC degradation by the UV/PMS process (0.3 mM PMS). It is assumed that SO₄^•− radicals generated during UV/PMS are responsible for the superiority of this process compared to the UV/H₂O₂ and combined UV/H₂O₂/PMS processes. During the SR-AOPs, HO^• radicals can also be generated under elevated pH conditions by the reaction between SO₄^•− and hydroxyl ions. During the treatment of natural water by UV/PMS under the above experimental conditions, both SO₄^•− and HO^• radicals coexisted, enabling the degradation synergism. Although HO^• and SO₄^•− possess similar redox potentials, their chemistry differs significantly. SO₄^•− tends to be more selective, primarily attacking the organics through the electron transfer pathway. In contrast, HO^• is nonselective, engaging in electrophilic addition to π bonds of unsaturated hydrocarbons and aromatics, H-abstraction in aliphatic hydrocarbons, and electron transfers in halogenated hydrocarbons. SO₄^•− demonstrates higher energy barriers and selectivity compared to HO^•, largely due to steric hindrance associated with its relatively larger molecular size [67]. On the other hand, it can be assumed that NOM in natural water showed a larger inhibition on the UV/H₂O₂ process and the combined UV/H₂O₂/PMS system than UV/PMS due to the lower rate constants of HO^• with NOM than SO₄^•− [58].

3.4. Evaluation of Water Quality Based on the Toxicity Assessment

Considering the often challenging task of detecting all potential transformation products (TPs), particularly in natural water rich in NOM, the utilization of toxicological tests emerges as a valuable strategy. Toxicological tests aid in evaluating the presence of potentially toxic TPs and the feasibility of certain treatments during drinking water preparation [57]. Figure 7 and Figure 8 present the results of toxicity assessment by Vibrio fischeri and Pseudomonas putida in ultrapure water and natural water before and after applying UV-based AOPs. Generally, before the application of the UV, UV/H₂O₂, UV/PMS, and UV/PMS/H₂O₂ processes, toxicity was measured in all water samples, and the inhibition on the tested species was negligible (<10%). However, it is necessary to note that an exception occurred during the evaluation of water quality after applying the UV/PMS/H₂O₂ process, where the inhibition was less than 2%. For this reason, the obtained results are not presented in Figure 7 and Figure 8. After the UV/PMS/H₂O₂ process was applied, the final TPs of the oxidative degradation of target contaminants and/or other constituents of the water matrices did not exhibit significant toxicity towards the Vibrio fischeri and Pseudomonas putida strain. In terms of techno-economic feasibility (oxidant consumption) and treated water quality, the UV/PMS/H₂O₂ process can be an adequate alternative to both the conventional UV/H₂O₂ treatment and the newer UV/PMS process for the degradation of target contaminants in water.

The highest inhibition rates were recorded during UV photolysis (600 mJ/cm²), reaching 50% and 35% inhibition to V. fisheri and P. putida, respectively. The increase from the initial inhibition indicates the potential formation of more toxic TPs during the UV treatment, which are further degraded, resulting in a reduction in toxicity to 25% by V. fischeri, i.e., to 18% by P. putida (1400 mJ/cm²). When applying both toxicity assays, in the UV/H₂O₂ and UV/PMS systems, the toxicity level was lower (20%) and decreased faster under lower UV fluence than during the UV treatment alone. The obtained results can be attributed to the faster decomposition of generated oxidative TPs during these oxidation processes compared to direct UV photolysis, which is also confirmed by the higher rate of decomposition of target contaminants when using the UV-based AOPs (Section 3.1).

In natural water samples, a slightly higher inhibition was obtained using the V. fischeri, suggesting that this test was more sensitive compared to P. putida (Figure 8). By applying the V. fischeri test, the inhibition reached 40% during UV photolysis and UV-driven AOPs (0.03 mM H₂O₂ and PMS), and the toxicity dropped to approximately 20% with an increase in the UV fluence. However, a higher initial concentration of H₂O₂ of 0.3 mM resulted in a significant increase in toxicity (70% in the intermediate stage of treatment), which indicates the fact that more toxic intermediates are formed compared to UV/PMS (45%). The reason for the sudden increase in toxicity, which then decreased to 30%, may also be a consequence of the oxidative transformation of NOM present in natural water, whereby more polar organic compounds can be formed that might exhibit toxicity towards the V. fischeri strain. The highest degree of toxicity to P. putida was observed at a UV fluence of 600 mJ/cm² (30%), which further decreased to 8–12% (1400 mJ/cm²), probably as a result of further oxidative degradation of toxic intermediates.

Several authors have noted that AOPs at certain stages of treatment can result in an increase in water toxicity. This increase is often attributed to undetected TPs, which are primarily responsible for the observed rise in toxicity [68]. A similar development of toxicity as in our research (an increasing trend in the intermediate stages of treatment, which later decreases) during UV/H₂O₂ treatment was also indicated in other relevant studies based on the formation and successive decomposition of oxidation products, which exhibit a higher degree of inhibition than the parent compound [69]. The potential toxicity of 4-MBC and EHMC and the by-products that may arise from oxidative degradation using UV-based AOPs have not been thoroughly investigated. A study by Hsieh et al. [40] investigated toxicity using the Microtox^® acute toxicity test (V. fischeri) after the application of the UV/persulfate process for the degradation of 4-MBC in water. The results indicate that the toxicity continuously increased during the first 20 min. of UV/persulfate treatment and then decreased after 30 min. They assumed that additional unidentified transformation products of 4-MBC were accountable for the rise in toxicity. In addition, the authors pointed out that before implementing UV/persulfate treatment in practical applications, it is crucial to carefully conduct investigations into the source of toxicity and perform ecotoxicological assessments.

3.5. Removal of Sunscreen Agents by Combined UV-Driven AOPs and Adsorption on Magnetic Biochar

Magnetic biochar was utilized after the UV-driven treatments of natural water were investigated. The specific surface area of the pristine biochar was 196.23 m²/g, with micro and mesopore volumes of 0.116 cm³/g and 0.034 cm³/g, respectively. Following modification with magnetic nanoparticles, the specific surface area and micropore volume of the biochar decreased to 116.31 m²/g and 0.071 cm³/g, respectively, while the mesopore volume increased to 0.260 cm³/g. This indicates that the magnetic nanoparticles accumulated on the biochar surface during the precipitation process, resulting in the formation of a mesoporous structure [70]. Similar observations were reported by Chen et al. [71], who also noted a decrease in the specific surface area of biochar after modification with Fe₃O₄. Additionally, the average pore diameters of BC and MBC were measured at 3.01 nm and 9.76 nm, respectively, further confirming the mesoporous structure of these materials (2–50 nm), as defined by IUPAC classification [72]. In the XRD pattern of biochar, a peak at 22.58° was assigned to a crystallographic plane of cellulose, while another peak at 29.2° was observed, attributed to the (104) plane of calcite (CaCO₃), indicating the presence of calcium carbonate [73,74]. Additionally, two broad peaks centered at 2θ = 23.1 and 43° corresponding to the (002) and (100) planes of graphitic carbon materials, respectively, were noted. The absence of sharp peaks in the XRD pattern of BC indicates the lack of a discernible crystalline structure, suggesting that the biochar structures are entirely amorphous [75,76]. After the modification of biochar with magnetite nanoparticles, crystal planes corresponding to (111), (220), (311), (222), (400), (422), (333), and (440) with respective 2θ values of 18.4°, 30.2°, 35.7°, 37.1°, 43.4°, 53.6°, 57.2°, and 63.1° were matched with those of magnetite, suggesting the successful coating of the biochar materials with Fe₃O₄ nanoparticles [70,77]. The phase identification and crystalline structures of magnetic biochar were assessed through XRD analysis. Figure 9a illustrates the diffraction patterns of both unmodified and modified biochar materials. The FTIR spectra of both the unmodified and magnetic biochar are presented in Figure 9b. In both spectra, the peaks observed at 3430 cm⁻¹ and 1630 cm⁻¹ correspond to the stretching and bending vibrations of the -OH groups from physically adsorbed water molecules. Additionally, in the unmodified biochar, the peaks observed at 1441 cm⁻¹ were attributed to aliphatic C–H groups and C=C stretching in aromatic rings, and the peaks at 882 cm¹ were attributed to aromatic C–H bending vibration [78]. After the modification of biochar with magnetite nanoparticles, new peaks appeared at 630 cm⁻¹ and 580 cm⁻¹. These peaks correspond to the stretching vibrations of the Fe–O bonds in the crystalline lattice of magnetite, confirming the presence of magnetite nanoparticles as well as the successful of coating of the biochar [70,79].

UV-driven AOPs accompanied by adsorption on magnetic biochar were elucidated in order to remove residual UV filters and possible transformation products from the water. As mentioned in Section 3.3, using a lower concentration of H₂O₂ and PMS (0.03 mM) in natural water enables approximately 79% degradation of UV filters in UV/H₂O₂ and UV/PMS, and 74% degradation in the combined UV/PMS/H₂O₂ process at both oxidants concentration of 0.015 mM. To remove residual 4-MBC and EHMC, potentially present transformation products, and other constituents in natural water, adsorption on magnetic biochar was applied for the first time after the UV-based AOPs. The adsorption of 4-MBC and EHMC on magnetic biochar was examined over 240 min and the obtained results are presented as changes in the concentration of 4-MBC and EHMC (C_t/C_o) as a function of contact time (Figure 10). The results revealed that the adsorption equilibrium of UV filters on magnetic biochar was achieved after 30 min of contact with this adsorbent.

The adsorption mechanism underlying 4-MBC and EHMC on magnetic biochar can be explained by considering the surface charge of this material (pHpzc), as well as the pKa values of 4-MBC and EHMC. The pHpzc represents the pH at which the surface of the biochar carries a neutral charge, indicating an equilibrium between the adsorption of H⁺ and OH⁻ ions [80]. The pHpzc of magnetic biochar in this work was determined to be 6.7, suggesting that under the investigated conditions (pH = 7.8), the surface charge of this material is predominantly negative. On the other hand, the pKa values of 4-MBC and EHMC are pKa = 8 and pKa = 9, respectively, indicating that these pollutants, under the experimental pH, are primarily in their protonated form. Hence, electrostatic interactions between these pollutants and the magnetic biochar surface are established, significantly contributing to their removal from water. In addition to electrostatic interactions, the surface of magnetic biochar contains a large number of oxygen-containing functional groups such as carboxylic, phenolic, and lactonic acids, as well as magnetic Fe-O, etc. [81]. Therefore, other interactions such as hydrogen bonding and π–π stacking are also involved in the mechanism of removal of 4-MBC and EHMC using magnetic biochar. Specifically, the functional groups on the surface of magnetic biochar can form strong hydrogen bonds with C=O and -O- on the molecular structure of 4-MBC and EHMC, respectively, while simultaneously acting as electron acceptors, forming π–π electron interactions with 4-MBC and EHMC. Additionally, 4-MBC and EHMC can effectively bind with the hydrophobic sites on the surface of magnetic biochar, thus facilitating hydrophobic interactions, which are important driving forces for their absorption onto the magnetic biochar.

The given results imply that the coupling of AOPs and adsorption not only facilitates the removal of residual target contaminants but can also assist in the elimination of oxidation by-products and other compounds found in the natural matrix. Since the application of magnetic biochar provides easy separation from water, regeneration, and reuse, the combined approach presents a promising and sustainable alternative in water treatment. Moreover, biochar, as a green technology, offers the advantage of being derived from renewable biomass sources, making it an environmentally friendly option for pollutant removal.

4. Conclusions

This study indicates the significant potential of applying UV-driven AOPs in water treatment for the degradation of organic UV filters. Based on the experimental results, the following main conclusions were derived:

• i.. Compared to UV photolysis alone, the addition of PMS and/or H₂O₂ significantly improves the degradation of UV filters. The existence of the synergy of oxidants PMS and H₂O₂ with UV activation significantly improves the degradation of UV filters in ultrapure water.

• ii.. Specific quenching tests revealed that HO^• played a key role in the UV/H₂O₂ process. HO^• and SO₄^•− radicals were present jointly in the UV/PMS and UV/PMS/H₂O₂ systems, while SO₄^•− predominantly contributed to the degradation in UV/PMS.

• iii.. The potential and benefits of the UV/PMS application were notable in the treatment of UV filters in NOM-reach natural water, exhibiting lower and finally negligible toxicity of treated water to V. fischeri and P. putida strains.

• iv.. In natural water, EHMC degrades faster than 4-MBC, with the highest degradation rate obtained during UV/PMS due to more selective oxidation by SO₄^•− radicals.

• v.. Combining UV-driven AOPs with magnetic biochar adsorption proved to be an efficient and sustainable method for completely removing UV filters, TPs, and other water matrix constituents. This approach not only ensures effective pollutant removal but also offers an environmentally friendly and sustainable solution.

Author Contributions

Conceptualization: J.M.J., T.S., J.A. and A.T.; methodology: J.M.J., T.S., J.A. and A.T.; formal analysis: T.S., M.K, J.N., and D.T.; investigation: T.S., M.K., J.N. and D.T.; data curation: T.S. and M.K.; writing—original draft preparation: T.S., J.N. and J.M.J.; writing—review and editing: J.M.J. and J.A.; visualization: T.S. and J.N. All authors have read and agreed to the published version of the manuscript.

Data Availability Statement

Data presented in this study are available upon request and are available at Zenodo (https://doi.org/10.5281/zenodo.11366731) URL (accessed on 28 May 2024).

Acknowledgments

The authors gratefully acknowledge the financial support of the Ministry of Science, Technological Development and Innovation of the Republic of Serbia (Grants No. 451-03-66/2024-03/200125 and 451-03-65/2024-03/200125) and European Union (Grant agreement No. 101059546, TwinSubDyn project).

Conflicts of Interest

The authors declare no conflict of interest.

Footnotes

Disclaimer/Publisher’s Note: The statements, opinions and data contained in all publications are solely those of the individual author(s) and contributor(s) and not of MDPI and/or the editor(s). MDPI and/or the editor(s) disclaim responsibility for any injury to people or property resulting from any ideas, methods, instructions or products referred to in the content.

Figures and Tables

Figure 1. (a,b) Degradation of sunscreen agents in ultrapure water using different processes: H2O2, PMS, PMS/H2O2, UV-C alone (control processes), and UV-driven AOPs. Experimental conditions: [4-MBC]0 = [EHMC]0 = 100 µg/L; [H2O2]0 = [PMS]0 = 0.03 mM (UV/H2O2 and UV/PMS); [H2O2]0 = [PMS]0 = 0.015 mM (UV/PMS/H2O2); pH = 7.0 ± 0.3.

Figure 2. 4-MBC and EHMC degradation rate constants in ultrapure water obtained during the UV-C photolysis, UV/H2O2, UV/PMS, and UV/PMS/H2O2 processes.

Figure 3. The quenching effect of MeOH and TBA on the degradation of 4-MBC in ultrapure water using the (a) UV/H2O2, (b) UV/PMS, and (c) UV/PMS/H2O2 processes. Experimental conditions: [4-MBC]0 = [EHMC]0 = 100 µg/L, [H2O2]0 = [PMS]0 = 0.03 mM, [MeOH]0 = [TBA]0 = 50 mM (UV/H2O2 and UV/PMS); [H2O2]0 = [PMS]0 = 0.015 mM, [MeOH]0 = [TBA]0 = 50 mM and 200 mM (UV/PMS/H2O2); pH = 7.0 ± 0.3.

Figure 4. The quenching effect of MeOH and TBA on the degradation of EHMC in ultrapure water using the (a) UV/H2O2, (b) UV/PMS, and (c) UV/PMS/H2O2 processes. Experimental conditions: [4-MBC]0 = [EHMC]0 = 100 µg/L, [H2O2]0 = [PMS]0 = 0.03 mM, [MeOH]0 = [TBA]0 = 50 mM (UV/H2O2 and UV/PMS); [H2O2]0 = [PMS]0 = 0.015 mM, [MeOH]0 = [TBA]0 = 50 mM and 200 mM (UV/PMS/H2O2); pH = 7.0 ± 0.3.

Figure 5. Degradation of sunscreen agents in natural water using different processes: UV and UV-driven AOPs (a,b). Experimental conditions: [4-MBC]0 = [EHMC]0 = 100 µg/L; [H2O2]0 = [PMS]0 = 0.03 and 0.3 mM (UV/H2O2 and UV/PMS); [H2O2]0 = [PMS]0 = 0.015 mM (UV/PMS/H2O2); pH = 7.81 ± 0.23.

Figure 6. 4-MBC and EHMC degradation rate constants in natural water obtained during the UV-C photolysis and UV-driven AOPs (a,b). Experimental conditions: [4-MBC]0 = [EHMC]0 = 100 µg/L; [H2O2]0 = [PMS]0 = 0.03 and 0.3 mM (UV/H2O2 and UV/PMS); [H2O2]0 = [PMS]0 = 0.015 mM (UV/PMS/H2O2); pH = 7.81 ± 0.23.

Figure 7. Toxicity evaluation by (a) Vibrio fischeri and (b) Pseudomonas putida during 4-MBC and EHMC degradation in ultrapure water.

Figure 8. Toxicity evaluation by (a) Vibrio fischeri and (b) Pseudomonas putida during 4-MBC and EHMC degradation in natural water.

Figure 9. (a) XRD pattern and (b) FTIR spectra of pristine biochar (BC) and magnetic biochar (MBC).

Figure 10. Adsorption of 4-MBC and EHMC on magnetic biochar after the application of (a) UV/H2O2, (b) UV/PMS, and (c) UV/PMS/H2O2 in natural water. Experimental conditions: [H2O2]0 = [PMS]0 = 0.03 (UV/H2O2 and UV/PMS); [H2O2]0 = [PMS]0 = 0.015 mM (UV/PMS/H2O2); [adsorbent]0 = 0.3 g/L; reaction time 30–240 min; pH = 7.81 ± 0.23.

Table 1

Literature review on the degradation of 4-MBC and EHMC in water by AOPs.

Compound.	Matrix	Applied Process	Conclusions	Reference
A mixture of eight UV filters, including 4-MBC and EHMC	Deionized water,drinking water, and wastewater effluent	Ozonation and UV254nm/H2O2	The ozone-refractory UV filters required ˃1 mg O3/mg DOC to achieve over 80% removal. The elimination of UV filters by UV/H2O2 was primarily controlled by HO•, with minimal contribution from UV photolysis.	[9]
A mixture of 21 UV filters, including 4-MBC and EHMC	Ultrapure water, river, sea, swimming pool water	UV365nm-A and UV254nm-C photolysis; UV-A/TiO2 photocatalysis and UV-C/H2O2	UV-C showed a high removal efficiency compared to UV-A radiation. The degradation yield was dependent on the water matrix. The UV-C/H2O2 system was successfully employed, achieving over 90% degradation of UV filters.	[10]
EHMC	Ultrapure water	UV200–400nm photolysis	EHMC was not fully mineralized, forming non-persistent photo-transformation products that later degraded.	[37]
4-MBC	Wastewater	Fenton,UV light irradiation,photo-Fenton	The photo-Fenton process exhibited superior degradation efficiency for 4-MBC compared to the Fenton process, achieving rates of 97% and 66%, respectively. Under UV irradiation, 4-MBC concentration decreased by 85%.	[38]
4-MBC	Contaminated water	Photolytic and photocatalytic degradation	UV filter photolysis in natural water is initially slow but can be accelerated by photosensitizers and nutrients. The efficiency processes are influenced by the radiation source, catalyst type, and oxygen content.	[39]
4-MBC	Deionized water, swimming pool water (SPW)	UV-activated persulfate	Results show that UV/persulfate effectively removes 4-MBC, with a kobs of 0.1349 min−1 under the specified conditions. When applied to SPW, 4-MBC removal notably decreases. It is recommended to pretreat the water to remove coexisting substances such as organics and inorganics to enhance the performance of the UV/persulfate system.	[40]

Table 2

The physicochemical properties of 4-MBC and EHMC.

Compound	Information		Chemical Structure *
4-MBC 1	Molecular formula: Molecular weight: Density:logKow:Boiling point:Solubility in water:Vapor pressure:	C18H22O254.37 g/mol1.064 ± 0.06 g/cm35.92357 °C1.3 mg/L at 20 °C1.0 × 10−3 Pa at 25 °C	[Image omitted. Please see PDF.]
EHMC 2	Molecular formula:Molecular weight: Density:logKow:Boiling point:Solubility in water:Vapor pressure:	C18H26O3290.40 g/mol1.005 ± 0.07 g/cm36.01193 °C0.2 mg/L at 20 °C1.38 × 10−5 Pa at 25 °C	[Image omitted. Please see PDF.]

* Data from ACD/ChemSketch; ¹ reference [48]; ² reference [49].

References

1. Baldermann, C.; Loren, S. UV-Strahlung in Deutschland: Einflüsse des Ozonabbaus unddes Klimawandels sowie Maßnahmen zum Schutz der Bevölkerung. Bundesgesundheitsblatt-Gesundheitsforschung-Gesundheitsschutz; 2019; 62, pp. 639-645. [DOI: https://dx.doi.org/10.1007/s00103-019-02934-w]

2. Chavda, V.P.; Acharya, D.; Hala, V.; Daware, S.; Vora, L.K. Sunscreens: A comprehensive review with the application of nanotechnology. J. Drug Deliv. Sci. Technol.; 2023; 86, 104720. [DOI: https://dx.doi.org/10.1016/j.jddst.2023.104720]

3. Zhang, S.; Wang, T. Preparation of enzymolysis porous corn starch composite microcapsules embedding organic sunscreen agents and its UV protection performance and stability. Carbohydr. Polym.; 2023; 314, 120903. [DOI: https://dx.doi.org/10.1016/j.carbpol.2023.120903]

4. Cadena-Aizaga, M.I.; Montesdeoca-Esponda, S.; Torres-Padrón, M.E.; Sosa-Ferrera, Z.; Santana-Rodríguez, J.J. Organic UV filters in marine environments: An update of analytical methodologies, occurrence and distribution. Trends Environ. Anal. Chem.; 2020; 25, e00079. [DOI: https://dx.doi.org/10.1016/j.teac.2019.e00079]

5. Duis, K.; Junker, T.; Coors, A. Review of the environmental fate and effects of two UV filter substances used in cosmetic products. Sci. Total Environ.; 2022; 808, 151931. [DOI: https://dx.doi.org/10.1016/j.scitotenv.2021.151931]

6. Carve, M.; Nugegoda, D.; Allinson, G.; Shimeta, J. A systematic review and ecological risk assessment for organic ultraviolet filters in aquatic environments. Environ. Pollut.; 2021; 268, 115894. [DOI: https://dx.doi.org/10.1016/j.envpol.2020.115894]

7. Gomez-Regalado, M.C.; Martín-Pozo, L.; Cantarero-Malagόn, S.; Hidalgo, F.; Gόmez, A.Z. Ultra-high performance liquid chromatography tandem mass spectrometry analysis of UV filters in marine mussels (Mytilus galloprovinciallis) from the southern coast of Spain. Microchem. J.; 2021; 171, 106800. [DOI: https://dx.doi.org/10.1016/j.microc.2021.106800]

8. Carvalho, P.G.; Naveira, R.I.; Granone, L.I.; Mendive, C.B.; Massa, A.E.; Churio, S. A comparative review of natural and synthetic UV filters: Gadusol and benzophenone-3 as representative examples. Environ. Adv.; 2023; 13, 100404. [DOI: https://dx.doi.org/10.1016/j.envadv.2023.100404]

9. Seo, C.; Shinc, J.; Lee, M.; Lee, W.; Yoom, H.; Sona, H.; Jang, S.; Lee, Y. Elimination efficiency of organic UVfilters during ozonation and UV/H₂O₂ treatment of drinking water and wastewater effluent. Chemosphere; 2019; 230, pp. 248-257. [DOI: https://dx.doi.org/10.1016/j.chemosphere.2019.05.028]

10. Celerio, M.; Facorro, R.; Dagnac, T.; Vilar, V.J.P.; Liompart, M. Photodegradation behavior of multiclass ultraviolet filters in the aquatic environment: Removal strategies and photoproduct identification by liquid chromatography-high resolution mass spectrometry. J. Chromatogr. A; 2019; 1596, pp. 8-19. [DOI: https://dx.doi.org/10.1016/j.chroma.2019.02.065]

11. Conway, A.J.; Gonsior, M.; Clark, C.; Heyes, A.; Mitchelmore, C.L. Acute toxicity of the UV filter oxybenzone to the coral Galaxea fascicularis. Sci. Total Environ.; 2021; 796, 148666. [DOI: https://dx.doi.org/10.1016/j.scitotenv.2021.148666]

12. Lopez-Arago, N.; Dominguez, A.; Munoz, M.; De Pedro, Z.M.; Casas, J.A. Application of catalytic wet peroxide oxidation for sunscreen agents breakdown. J. Environ. Chem. Eng.; 2024; 12, 112410. [DOI: https://dx.doi.org/10.1016/j.jece.2024.112410]

13. Sun, J.; Rene, E.R.; Tao, D.; Lu, Y.; Jin, Q.; Lam, J.C.-H.; Leung, K.M.Y.; He, Y. Degradation of organic UV filters in the water environment: A concise review on the mechanism, toxicity, and tehnologies. J. Hazard. Mater.; 2024; 463, 132822. [DOI: https://dx.doi.org/10.1016/j.jhazmat.2023.132822]

14. Guimarães, T.; da Silva, E.M.G.; de Aguiar, A.C.M.; da Costa, K.M.; Mielke, K.C.; Mendes, K.F.; da Silva, A.A.; de Carvalho Teixeira, A.P.; Lopes Moreira, R.P. Exploration of Eco-Friendly Hydrochar’s Potential in Advanced Oxidative Processes for Dicamba Degradation within a Circular Bio-Economy Framework. Processes; 2023; 11, 3244. [DOI: https://dx.doi.org/10.3390/pr11113244]

15. Néstor, A.; Urbina-Suarez, N.A.; López-Barrera, G.L.; García-Martínez, J.B.; Barajas-Solano, A.F.; Machuca-Martínez, F.; Zuorro, A. Enhanced UV/H₂O₂ System for the Oxidation of Organic Contaminants and Ammonia Transformation from Tannery Effluents. Processes; 2023; 11, 3091. [DOI: https://dx.doi.org/10.3390/pr11113091]

16. Zhou, S.; Jiang, L.; Wang, H.; Yang, J.; Yuan, X.; Wang, H.; Liang, J.; Li, X.; Li, H.; Bu, Y. Oxygen Vacancies Modified TiO₂/O-Terminated Ti₃C₂ Composites: Unravelling the Dual Effects between Oxygen Vacancy and High-Work-Function Titanium Carbide. Adv. Funct. Mater; 2023; 33, 2307702. [DOI: https://dx.doi.org/10.1002/adfm.202307702]

17. Afitiri, A.-R.; Aram, S.A.; Martienssen, M. Systematic review of the effects of advanced oxidation processes integration with solar water disinfection for improved drinking water production. Waste. Manag. Bull.; 2024; 1, pp. 52-59. [DOI: https://dx.doi.org/10.1016/j.wmb.2023.08.005]

18. Ramasundaram, S.; Balasanka, A.; Arokiyaraj, S.; Sumathi, P.; Oh, T.H. Multi-usable titanium dioxide and poly(vinylidene fluoride) composite foam photocatalyst for degradation of organic pollutants. Appl. Surf. Sci.; 2023; 609, 155264. [DOI: https://dx.doi.org/10.1016/j.apsusc.2022.155264]

19. Singh, A.; Majumder, A.; Saidulu, D.; Bhattcharya, A.; Bhatnagar, A.; Gupta, A.K. Oxidative treatment of micropollutants present in wastewater: A special emphasis on transformation products, their toxicity, detection, and field-scale investigations. J. Environ. Manag.; 2024; 354, 120339. [DOI: https://dx.doi.org/10.1016/j.jenvman.2024.120339]

20. Xie, Z.-H.; He, C.-S.; Pei, D.-N.; Dong, Y.; Yang, S.-R.; Xiong, Z.; Zhou, P.; Pan, Z.-C.; Yao, G.; Lai, B. Review of characteristics, generation pathways and detection methods of singlet oxygen generated in advanced oxidation processes (AOPs). Chem. Eng. J.; 2023; 468, 143778. [DOI: https://dx.doi.org/10.1016/j.cej.2023.143778]

21. Wang, K.; Li, H.; Yu, W.; Ma, T. Insights into structural and functional regulation of chalcopyrite and enhanced mechanism of reactive oxygen species (ROS) generation in advanced oxidation process (AOP): A review. Sci. Total Environ.; 2024; 919, 170530. [DOI: https://dx.doi.org/10.1016/j.scitotenv.2024.170530]

22. Huang, Y.; Jia, Y.; Zuo, L.; Huo, Y.; Zhang, Y.; Hou, L. Comparison of VUV/H₂O₂ and VUV/PMS (peroxymonosulfate) for the degradation of unsymmetrical dimethylhydrazine in water. J. Water. Process. Eng.; 2022; 49, 102970. [DOI: https://dx.doi.org/10.1016/j.jwpe.2022.102970]

23. Álvarez, M.A.; Ruidíaz-Martínez, M.; Cruz-Quesada, G.; Victoria López-Ramón, M.; Rivera-Utrilla, J.; Sánchez-Polo, M.; Mota, A.J. Removal of parabens from water by UV-driven advanced oxidation processes. Chem. Eng. J.; 2020; 379, 122334. [DOI: https://dx.doi.org/10.1016/j.cej.2019.122334]

24. Tian, F.-X.; Ye, W.-K.; Xu, B.; Hu, X.-J.; Ma, S.-X.; Lai, F.; Gao, J.-Q.; Xing, H.-B.; Xia, W.-H.; Wang, B. Comparison of UV-induced AOPs (UV/Cl₂, UV/NH₂Cl, UV/ClO₂ and UV/H₂O₂) in the degradation of iopamidol: Kinetics, energy requirements and DBPs-related toxicity in sequential disinfection processes. Chem. Eng. J.; 2020; 398, 125570. [DOI: https://dx.doi.org/10.1016/j.cej.2020.125570]

25. Zhang, X.; Kamali, M.; Uleners, T.; Symus, J.; Zhang, S.; Liu, Z.; Costa, M.E.V.; Appels, L.; Cabooter, D.; Dewil, R. UV/TiO₂/periodate system for the degradation of organic pollutants–Kinetics, mechanisms and toxicity study. Chem. Eng. J.; 2022; 449, 137680. [DOI: https://dx.doi.org/10.1016/j.cej.2022.137680]

26. Wang, X.; Tang, W.; Jiang, L.; Feng, J.; Yang, J.; Zhou, S.; Li, W.; Yuan, X.; Wang, H.; Wan, J. et al. Mechanism insights into visible light-induced crystalline carbon nitride activating periodate for highly efficient ciprofloxacin removal. Chem. Eng. J.; 2023; 471, 144521. [DOI: https://dx.doi.org/10.1016/j.cej.2023.144521]

27. Jiang, L.; Li, W.; Wang, H.; Yang, J.; Chen, H.; Wang, X.; Yuan, X.; Wang, H. Non-radical activation of low additive periodate by carbon-doped boron nitride for acetaminophen degradation: Significance of high-potential metastable intermediates. J. Hazard. Mater.; 2024; 469, 133806. [DOI: https://dx.doi.org/10.1016/j.jhazmat.2024.133806]

28. Wang, X.; Huang, H.; Li, W.; Chen, H.; Liu, W.; Yang, J.; Yuan, X.; Wang, H.; Chen, Y.; Jiang, L. Ultra-thin carbon nitride activated periodate for efficient tetracycline degradation with the assistance of visible light. Sep. Purif. Technol.; 2024; 338, 126505. [DOI: https://dx.doi.org/10.1016/j.seppur.2024.126505]

29. Preethi, M.; Shanmugavel, S.P.; Kumar, G.; Yogalakshmi, K.N.; Gunasekaran, M.; Rajesh Banu, J. Recent progress in mineralization of emerging contaminants by advanced oxidation process: A review. Environ. Pollut.; 2024; 341, 122842. [DOI: https://dx.doi.org/10.1016/j.envpol.2023.122842]

30. Cheng, N.; Wang, B.; Wu, P.; Lee, X.; Xing, Y.; Chen, M.; Gao, B. Adsorption of emerging contaminants from water and wastewater by modified biochar: A review. Environ. Pollut.; 2021; 273, 116448. [DOI: https://dx.doi.org/10.1016/j.envpol.2021.116448]

31. Jagadeesh, N.; Sundaram, B. Adsorption of Pollutants from Wastewater by Biochar: A Review. J. Hazard. Mater. Adv.; 2023; 9, 100226. [DOI: https://dx.doi.org/10.1016/j.hazadv.2022.100226]

32. Zeghioud, H.; Fryda, L.; Djelal, H.; Assadi, A.; Kane, A. A comprehensive review of biochar in removal of organic pollutants from wastewater: Characterization, toxicity, activation/functionalization and influencing treatment factors. J. Water Process. Eng.; 2022; 47, 102801. [DOI: https://dx.doi.org/10.1016/j.jwpe.2022.102801]

33. Wang, Y.; Chen, L.; Zhu, Y.; Fang, W.; Tan, Y.; He, Z.; Liao, H. Research status, trends, and mechanisms of biochar adsorption for wastewater treatment: A scientometric review. Environ. Sci. Eur.; 2024; 36, 25. [DOI: https://dx.doi.org/10.1186/s12302-024-00859-z]

34. Yi, Y.; Huang, Z.; Lu, B.; Xian, J.; Pokeung Tsang, E.; Cheng, W.; Fang, J.; Fang, Z. Magnetic biochar for environmental remediation: A review. Bioresour. Technol.; 2020; 98, 122468. [DOI: https://dx.doi.org/10.1016/j.biortech.2019.122468]

35. Liang, L.; Xi, F.; Tan, W.; Meng, X.; Hu, B.; Wang, X. Review of organic and inorganic pollutants removal by biochar and biochar-based composites. Biochar; 2021; 3, pp. 255-281. [DOI: https://dx.doi.org/10.1007/s42773-021-00101-6]

36. Yin, Y.B.; Yang, S.Q.; Jia, Z.Y.; Zhang, H.; Gao, Y.; Zhang, X.C.; Zhong, H.J.; Zhou, Z.Q.; Zhang, X.; Zhou, H.F. Magnetic biochar based on furfural residue as an excellent candidate for efficient adsorption of Tetracycline, Bisphenol A, Congo red, and Cr⁶⁺. Environ. Sci. Pollut. Res.; 2023; 30, pp. 26510-26522. [DOI: https://dx.doi.org/10.1007/s11356-022-23978-7]

37. Jentzsch, F.; Olsson, O.; Westphal, J.; Reich, M.; Leder, C.; Kümmerer, K. Photodegradation of the UV filter ethylhexyl methoxycinnamate under ultraviolet light: Identification and in silico assessment of photo-transformation products in the context of grey water reuse. Sci. Total Environ.; 2016; 572, pp. 1092-1100. [DOI: https://dx.doi.org/10.1016/j.scitotenv.2016.08.017]

38. Ji, C.; Wang, C.; Sun, J.; Yu, Y. Degradation of 4-methylbenzylidene camphor (4-MBC) Using Fenton, UV Light Irradiation and Photo-Fenton. MATEC. Web. Conf.; 2017; 100, 03006. [DOI: https://dx.doi.org/10.1051/matecconf/201710003006]

39. Ahmed, M.B.; Johir, M.A.H.; Zhou, J.L.; Ngo, H.H.; Guo, W.; Sornalingam, K. Photolytic and photocatalytic degradation of organic UV filters in contaminated water. Curr. Opin. Green Sustain. Chem.; 2017; 6, pp. 85-92. [DOI: https://dx.doi.org/10.1016/j.cogsc.2017.06.010]

40. Hsieh, S.-C.; Lai, W.W.-P.; Lin, A.Y.-C. Kinetics and mechanism of 4-methylbenzylidene camphor degradation by UV-activated persulfate oxidation. Environ. Sci. Pollut. Res. Int.; 2021; 28, pp. 18021-18034. [DOI: https://dx.doi.org/10.1007/s11356-020-11795-9]

41. Ðurkić, T.; Molnar Jazić, J.; Kragulj Isakovski, M.; Maletić, S.; Tubić, A.; Dalmacija, B.; Agbaba, J. Ultraviolet/hydrogen peroxide oxidative degradation of 1,2,3-trichlorobenzene: Influence of water matrix and toxicity assessment. Environ. Eng. Sci.; 2019; 36, pp. 947-957. [DOI: https://dx.doi.org/10.1089/ees.2019.0020]

42. ISO 10712:1995 Water Quality-Pseudomonas Putida Growth Inhibition Test (Pseudomonas Cell Multiplication Inhibition Test); ISO: Geneva, Switzerland, 1995.

43. ISO 11348:2007 Water Quality-Determination of the Inhibitory Effect of Water Samples on the Light Emission of Vibrio fischeri (Luminescent Bacteria Test); ISO: Geneva, Switzerland, 2007.

44. Díaz-Cruz, M.S.; Gago-Ferrero, P.; Llorca, M.; Barceló, D. Analysis of UV filters in tap water and other clean waters in Spain. Anal. Bioanal. Chem.; 2012; 402, pp. 2325-2333. [DOI: https://dx.doi.org/10.1007/s00216-011-5560-8]

45. Rascόn, A.J.; Rocío-Bautista, P.; Palacios-Colόn, L.; Ballesteros, E. Easy determination of benzophenone and its derivatives in sunscreen samples by direct-immersion solid-phase microextraction and gas chromatography–mass spectrometry. J. Pharm. Biomed. Anal.; 2023; 236, 115711. [DOI: https://dx.doi.org/10.1016/j.jpba.2023.115711]

46. APHA-AWWA-WEF. Standard Methods for the Examination of Water and Wastewater; 22nd ed. American Public Health Association/American Water Works Association/Water Environment Federation: Washington, DC, USA, 2012.

47. Šolić, M.; Maletić, S.; Kragulj Isakovski, M.; Nikić, J.; Watson, M.; Kónya, Z.; Tricković, J. Comparing the Adsorption Performance of Multiwalled Carbon Nanotubes Oxidized by Varying Degrees for Removal of Low Levels of Copper, Nickel and Chromium (VI) from Aqueous Solutions. Water; 2020; 12, 723. [DOI: https://dx.doi.org/10.3390/w12030723]

48. SCCS (Scientific Committee on Consumer Safety). Scientific Opinion on 4-Methylbenzylidene Camphor (4-MBC), Preliminary Version of 22 December, Final Version of 29 April 2022, SCCS/1640/21; SCCS: Luxembourg, 2022.

49. Gackowska, A.; Studziński, W.; Kudlek, E.; Dudziak, M.; Gaca, J. Estimation of physicochemical properties of 2-ethylhexyl-4-methoxycinnamate (EHMC) degradation products and their toxicological evaluation. Environ. Sci. Pollut. Res.; 2018; 25, pp. 16037-16049. [DOI: https://dx.doi.org/10.1007/s11356-018-1796-6]

50. Bracamontes-Ruelas, A.R.; Ordaz-Díaz, L.A.; Bailón-Salas, A.M.; Ríos-Saucedo, J.C.; Reyes-Vidal, Y.; Reynoso-Cueva, L. Emerging Pollutants in Wastewater, Advanced Oxidation Processes as an Alternative Treatment and Perspectives. Processes; 2022; 10, 1041. [DOI: https://dx.doi.org/10.3390/pr10051041]

51. Mergenbayeva, S.; Poulopoulos, S.G. Comparative Study on UV-AOPs for Efficient Continuous Flow Removal of 4-tert-Butylphenol. Processes; 2022; 10, 8. [DOI: https://dx.doi.org/10.3390/pr10010008]

52. Guerra-Rodríguez, S.; Rodríguez, E.; Rodríguez-Chueca, J. Pilot-scale regeneration of wastewater through intensified sulfate radical-based advanced oxidation processes (PMS/UV-A, PMS/H₂O₂/UV-A, and PMS/O₃): Inactivation of bacteria and mechanistic considerations. Chem. Eng. J.; 2023; 469, 143859. [DOI: https://dx.doi.org/10.1016/j.cej.2023.143859]

53. Amanollahi, H.; Moussavi, G.; Giannakis, S. Enhanced vacuum UV-based process (VUV/H₂O₂/PMS) for the effective removal of ammonia from water: Engineering configuration and mechanistic considerations. J. Hazard. Mater.; 2021; 402, 123789. [DOI: https://dx.doi.org/10.1016/j.jhazmat.2020.123789]

54. Zhong, D.; Zhou, Z.; Ma, W.; Ma, J.; Lv, W.; Feng, W.; Du, X.; He, F.u. Study on degradation of chloramphenicol by H₂O₂/PMS double-oxidation system catalyzed by pipe deposits from water networks. J. Environ. Chem. Eng.; 2022; 10, 107529. [DOI: https://dx.doi.org/10.1016/j.jece.2022.107529]

55. Sun, H.; Zhang, B.; Wang, N.a.; Zhang, N.; Ma, Y.; Zang, L.; Li, Z.; Xue, R. Refractory organics removal in PMS and H₂O₂/PMS oxidation system activated by biochar/ nZVI/MoS₂ composite: Synthesis, performance, mechanism and dosing methods. J. Environ. Chem. Eng.; 2023; 11, 109134. [DOI: https://dx.doi.org/10.1016/j.jece.2022.109134]

56. Molnar Jazić, J.; Đurkić, T.; Bašić, B.; Watson, M.; Apostolović, T.; Tubić, A.; Agbaba, J. Degradation of a chloroacetanilide herbicide in natural waters using UV activated hydrogen peroxide, persulfate and peroxymonosulfate processes. Environ. Sci. Water. Res. Technol.; 2020; 6, pp. 2800-2815. [DOI: https://dx.doi.org/10.1039/D0EW00358A]

57. Nitulescu, G.; Lupuliasa, D.; Adam-Dima, I.; Nitulescu, G.M. Ultraviolet Filters for Cosmetic Applications. Cosmetics; 2023; 10, 101. [DOI: https://dx.doi.org/10.3390/cosmetics10040101]

58. Guan, Y.-H.; Chen, J.; Chen, L.-J.; Jiang, X.-X.; Fu, Q. Comparison of UV/H₂O₂, UV/PMS, and UV/PDS in Destruction of Different Reactivity Compounds and Formation of Bromate and Chlorate. Front. Chem.; 2020; 8, 521198. [DOI: https://dx.doi.org/10.3389/fchem.2020.581198]

59. Du, J.; Wang, C.; Zhao, Z.; Cui, F.; Ou, Q.; Liu, J. Role of oxygen and superoxide radicals in promoting H₂O₂ production during VUV/UV radiation of water. Chem. Eng. Sci.; 2021; 241, 116683. [DOI: https://dx.doi.org/10.1016/j.ces.2021.116683]

60. Hoang, N.T.; Mwazighe, F.M. The various effects of hydrogen phosphate and bicarbonate in the degradation of some pollutants in the UV/chlorine and the UV/H₂O₂ processes. J. Water. Process. Eng.; 2024; 57, 104646. [DOI: https://dx.doi.org/10.1016/j.jwpe.2023.104646]

61. Ding, X.; Gutierrezb, L.; Croue, J.-P.; Lia, M.; Wang, L.; Wang, Y. Hydroxyl and sulfate radical-based oxidation of RhB dye in UV/H₂O₂ and UV/persulfate systems: Kinetics, mechanisms, and comparison. Chemosphere; 2020; 253, 126655. [DOI: https://dx.doi.org/10.1016/j.chemosphere.2020.126655]

62. Xiao, R.; He, L.; Luo, Z.; Spinney, R.; Wei, Z.; Dionysiou, D.D.; Zhao, F. An experimental and theoretical study on the degradation of clonidine by hydroxyl and sulfate radicals. Sci. Total Environ.; 2020; 710, 136333. [DOI: https://dx.doi.org/10.1016/j.scitotenv.2019.136333]

63. Liu, H.; Liu, F.; Zhang, J.; Zhou, J.; Bi, W.; Qin, J.; Hou, Q.; Ni, Y.; Shaozu, X.; Yang, C. Degradation of methyl orange by pyrite activated persulfate oxidation: Mechanism, pathway and influences of water substrates. Water Sci. Technol.; 2022; 85, pp. 2912-2927. [DOI: https://dx.doi.org/10.2166/wst.2022.134]

64. Behnami, A.; Aghayani, E.; Benis, K.Z.; Sattarie, M.; Pourakbar, M. Comparing the efficacy of various methods for sulfate radical generation for antibiotics degradation in synthetic wastewater: Degradation mechanism, kinetics study, and toxicity assessment. RSC Adv.; 2022; 12, pp. 14945-14956. [DOI: https://dx.doi.org/10.1039/D2RA01618D]

65. Lado Ribeiro, A.R.; Moreira, N.F.F.; Puma, G.L.; Silva, A.M.T. Impact of water matrix on the removal of micropollutants by advanced oxidation technologies. Chem. Eng. J.; 2019; 363, pp. 155-173. [DOI: https://dx.doi.org/10.1016/j.cej.2019.01.080]

66. Ratchnashree, S.R.; Karmegam, N.; Selvam, M.; Manikandan, S.; Deena, S.R.; Subbaiya, R.; Vickram, A.S.; Kim, W.; Govarthanan, M. Advanced technologies for the determination of quantitative structure-activity relationships and degradation efficiency of micropollutants and their removal in water—A review. Sci. Total Environ.; 2023; 904, 166563. [DOI: https://dx.doi.org/10.1016/j.scitotenv.2023.166563]

67. Scaria, J.; Nidheesh, P.V. Comparison of hydroxyl-radical-based advanced oxidation processes with sulfate radical-based advanced oxidation processes. Curr. Opin. Chem. Eng.; 2022; 36, 100830. [DOI: https://dx.doi.org/10.1016/j.coche.2022.100830]

68. Stankov, V.; Novak Stankov, M.; Cvetnić, M.; Sigurnjak Bureš, M.; Ukić, S.; Kučić Grgić, D.; Lončarić Božić, A.; Kušić, H.; Bolanča, T. Environmental aspects of UV-C-based processes for the treatment of oxytetracycline in water. Environ. Pollut.; 2021; 277, 116797. [DOI: https://dx.doi.org/10.1016/j.envpol.2021.116797]

69. Son, Y.; Lee, Y.-M.; Zoh, K.-D. Kinetics and degradation mechanism of tris (1-chloro-2-propyl) phosphate in the UV/H₂O₂ reaction. Chemosphere; 2020; 260, 127461. [DOI: https://dx.doi.org/10.1016/j.chemosphere.2020.127461]

70. Guo, D.; Wu, J.; Feng, D.; Zhang, Y.; Zhu, X.; Luo, Z.; Kang, Y.; Zhao, Y.; Sun, S. Mechanism of efficient magnetic biochar for typical aqueous organic contaminant combined-adsorption removal. Fuel Process. Technol.; 2023; 247, 107795. [DOI: https://dx.doi.org/10.1016/j.fuproc.2023.107795]

71. Chen, X.; Lin, J.; Su, Y.; Tang, S. One-Step Carbonization Synthesis of Magnetic Biochar with 3D Network Structure and Its Application in Organic Pollutant Control. Int. J. Mol. Sci.; 2022; 23, 12579. [DOI: https://dx.doi.org/10.3390/ijms232012579]

72. McNaught, A.D.; Wilkinson, A. IUPAC. Compendium of Chemical Terminology; 2nd ed. Blackwell Scientific Publications: Oxford, UK, 1997.

73. Stylianou, M.; Christou, A.; Dalias, P.; Polycarpou, P.; Michael, C.; Agapiou, A.; Papanastasiou, P.; Fatta-Kassinos, D. Physicochemical and structural characterization of biochar derived from the pyrolysis of biosolids, cattle manure and spent coffee grounds. J. Energy Inst.; 2020; 93, pp. 2063-2073. [DOI: https://dx.doi.org/10.1016/j.joei.2020.05.002]

74. Sahoo, S.S.; Vijay, V.K.; Chandra, R.; Kumar, H. Production and characterization of biochar produced from slow pyrolysis of pigeon pea stalk and bamboo. Clean Eng. Technol.; 2021; 3, 100101. [DOI: https://dx.doi.org/10.1016/j.clet.2021.100101]

75. Hu, B.; Yan, N.; Zheng, Z.; Xu, L.; Xie, H.; Chen, J. Recyclable Magnesium-Modified Biochar Beads for Efficient Removal of Phosphate from Wastewater. Nanomaterials; 2023; 13, 966. [DOI: https://dx.doi.org/10.3390/nano13060966]

76. Kim, D.-Y.; Jung, G.-B. Effects of Pyrolysis and Ball-Milling on the Physicochemical and Rhodamine B Removal Characteristics of Rice-Bran-Derived Biochar. Appl. Sci.; 2023; 13, 4288. [DOI: https://dx.doi.org/10.3390/app13074288]

77. Santhosh, C.; Daneshvar, E.; Tripathi, K.M.; Baltrėnas, P.; Kim, T.Y.; Baltrėnaitė, E.; Bhatnagar, A. Synthesis and characterization of magnetic biochar adsorbents for the removal of Cr(VI) and Acid orange 7 dye from aqueous solution. Environ. Sci. Pollut. Res.; 2020; 27, pp. 32874-32887. [DOI: https://dx.doi.org/10.1007/s11356-020-09275-1]

78. Zhang, C.; Zhang, N.; Xiao, Z.; Li, Z.; Zhang, D. Characterization of biochars derived from different materials and their effects on microbial dechlorination of pentachlorophenol in a consortium. RSC Adv.; 2019; 9, pp. 917-923. [DOI: https://dx.doi.org/10.1039/C8RA09410A]

79. Nikić, J.; Tubić, A.; Watson, M.; Maletić, S.; Šolić, M.; Majkić, T.; Agbaba, J. Arsenic Removal from Water by Green Synthesized Magnetic Nanoparticles. Water; 2019; 11, 2520. [DOI: https://dx.doi.org/10.3390/w11122520]

80. Murtaza, G.; Ahmed, Z.; Dai, D.-Q.; Iqbal, R.; Bawazeer, S.; Usman, M.; Rizwan, M.; Iqbal, J.; Akram, M.I.; Althubiani, A.S. et al. A review of mechanism and adsorption capacities of biocharbased engineered composites for removing aquatic pollutants from contaminated water. Front. Environ. Sci.; 2022; 10, 1035865. [DOI: https://dx.doi.org/10.3389/fenvs.2022.1035865]

81. Qu, J.; Shi, J.; Wang, Y.; Tong, H.; Zhu, Y.; Xu, L.; Wang, Y.; Zhang, B.; Tao, Y.; Dai, X. et al. Applications of functionalized magnetic biochar in environmental remediation: A review. J. Hazard. Mater.; 2022; 434, 128841. [DOI: https://dx.doi.org/10.1016/j.jhazmat.2022.128841]