# Removal of pharmaceuticals through UV-C/Performic acid advanced oxidation process: Kinetics and identification of reactive species

Christelle Nabintu Kajoka, Stephan Brosillon, Corine Reibel, Yacine Khadija Diop, Marcos Oliveira, Vincent Rocher, Ghassan Chebbo, Johnny Gasperi,

Julien Le Roux

► To cite this version:

Christelle Nabintu Kajoka, Stephan Brosillon, Corine Reibel, Yacine Khadija Diop, Marcos Oliveira, et al.. Removal of pharmaceuticals through UV-C/Performic acid advanced oxidation process: Kinetics and identification of reactive species. Journal of Hazardous Materials, 2025, 495, pp.139016. 10.1016/j.jhazmat.2025.139016. hal-05143664

HAL Id: hal-05143664

https://enpc.hal.science/hal-05143664v1

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# Removal of pharmaceuticals through UV-C/Performic acid advanced oxidation process: Kinetics and identification of reactive species

Christelle Nabintu Kajoka<sup>a</sup>, Stephan Brosillon<sup>b</sup>, Corine Reibel<sup>c</sup>, Yacine Khadija Diop<sup>a</sup>,
Marcos Oliveira<sup>d</sup>, Vincent Rocher<sup>d</sup>, Ghassan Chebbo<sup>a</sup>, Johnny Gasperi<sup>c</sup>, Julien Le Roux<sup>a,∗</sup>

## HIGHLIGHTS

* UV-C/performic acid (PFA) is an emerging efficient advanced oxidation process for the removal of pharmaceuticals.

* Singlet oxygen (¹O₂) is generated via PFA autodecomposition and further amplified by UV-C/PFA.

* Hydroxyl (HO•) and peroxyl (HCOO• and HCOOO•) radicals are the main reactive species generated under UV-C/ PFA.

* HO• radicals produced under UV-C/PFA are primarily generated by the background H₂O₂ concentration in the PFA solution.

## GRAPHICAL ABSTRACT

<figure><img src="image_2.png" /></figure>

## ARTICLE INFO

**Keywords:**

Peracid

Peracetic acid

Wastewater treatment

Transformation products

UV-C photolysis

## ABSTRACT

Performic acid (PFA), widely recognized for its disinfectant properties in wastewater, shows selective and limited reactivity in oxidizing micropollutants. This study investigates the activation of PFA through UV-C photolysis to generate an advanced oxidation process (UV-C/PFA) and enhance the degradation of six pharmaceuticals: lidocaine, furosemide, sulfamethoxazole, diclofenac, acetaminophen, and carbamazepine. The synergy of UV-C photolysis with PFA enhances the removal of PFA-persistent pharmaceuticals. For instance, diclofenac, acetaminophen, and sulfamethoxazole, initially unreactive with PFA, were entirely degraded within ten minutes under UV-C/PFA in a phosphate buffer solution. This increased reactivity results from generated reactive species like hydroxyl (HO•), peroxyl (R–O•) radicals and singlet oxygen (¹O₂), confirmed by electron paramagnetic resonance, with HO• primarily originating from the background H₂O₂ present in the PFA solution. While UV-C/ PFA produced fewer HO• than UV-C/H₂O₂, it has distinct advantages through the selective action of ¹O₂ and R–O• in degrading some pharmaceuticals. ¹O₂ was also detected in the PFA solution and could explain its selective reactivity, especially with compounds containing reduced sulfur or tertiary amine groups. Overall, UV-C/ PFA yields transformation products of lower molar mass compared to PFA, thus potentially increasing

<footnote><sup>a</sup> LEESU, ENPC, Institut Polytechnique de Paris, Univ Paris Est Creteil, Marne-la-Vallée, France</footnote>

<footnote><sup>b</sup> IEM - Institut Européen des Membranes, UMR 5635, Université de Montpellier, Montpellier F-34090, France</footnote>

<footnote><sup>c</sup> PAC Chimie Balard, Univ Montpellier, CNRS, ENSCM, Montpellier, France</footnote>

<footnote><sup>d</sup> SIAAP - Service Public de l'assainissement Francilien (SIAAP), Direction Innovation, Colombes F-92700, France</footnote>

<footnote><sup>e</sup> GERS-LEE - Laboratoire Eau et Environnement, Université Gustave Eiffel, Bouguenais F-44344, France</footnote>

<footnote>* Corresponding author.
E-mail address: julien.le-roux@u-pec.fr (J. Le Roux).</footnote>
mineralization. In wastewater effluent, UV-C/PFA improved pharmaceutical degradation, though scavenging effects by wastewater constituents reduced removal rates.

# 1. Introduction

The quality of surface water is increasingly deteriorating due to microbiological and chemical pollution from wastewater discharge. Conventional wastewater treatment plants (WWTPs) were not originally designed to efficiently remove this type of pollution, which has sparked interest in implementing additional (chemical or physical) treatments to improve the quality of discharged wastewater. Chlorine and ozone, while commonly used to remove microorganisms and micropollutants from water and wastewater, can also react with organic and inorganic compounds in water and generate disinfection by-products (DBPs). Some of these DBPs are known to be cytotoxic or genotoxic [50], and some can even be carcinogenic [42].

Peracids, specifically performic and peracetic acids (PFA and PAA, respectively), are gaining attention as alternative disinfectants because they form very few DBPs compared to traditional disinfectants [25,31, 39] and are easy to implement technically at existing WWTPs [34]. While PAA was the first peracid proposed for wastewater disinfection nearly 50 years ago [4,40], PFA is garnering increased interest due to its demonstrated superior effectiveness and lower operational costs compared to PAA in removing *Escherichia coli* and enterococci during wastewater treatment [39,53]. In addition, its effectiveness in disinfection has been proven against a broad spectrum of microorganisms, including bacteria, viruses, spores, antibiotic-resistant bacteria, mycobacteria, and microscopic fungi [10,22,31,47]. Although PFA is used as an alternative wastewater disinfectant at full scale in Europe [53,56], its ability to degrade organic micropollutants (OMPs) that are consistently present in wastewater and known to have adverse effects [6,63] has been scarcely investigated. For instance, Gagnon et al., [21] studied the degradation of eight pharmaceuticals (six parent compounds: clofibrate acid, ibuprofen, naproxen, triclosan, carbamazepine, diclofenac, and two metabolites: salicylic acid and 2-hydroxy-ibuprofen) from primary-treated wastewater by PFA and demonstrated poor reactivity (average removal rate < 8 %). Additionally, Ragazzo et al., [53] found that PFA had a weak oxidation power towards organic substances commonly found in wastewater, such as phenols and hormones like estrone. More recently, Nabintu Kajoka et al., [46] studied the removal of model organic compounds and pharmaceuticals by PFA, revealing the selective reactivity of PFA with reduced-sulfur moieties (total removal of ranitidine, 3-mercaptophenol, and benzenethiol in less than 1 min in a phosphate buffer solution) and deprotonated nitrogen-containing compounds (lidocaine, ranitidine) in both phosphate buffer solution and wastewater effluent. Based on the low reactivity of PFA with most pharmaceuticals, this study also demonstrated that PFA is generally a weaker oxidant compared to common oxidants such as ozone.

To enhance the removal rate of recalcitrant pharmaceuticals, UV-C photolysis can activate PFA to generate reactive species through a PFA-based advanced oxidation process (AOP). This approach, previously investigated for PAA (UV-C/PAA), resulted in an increased removal rate of organic compounds, attributed to the production of reactive radical species such as hydroxyl (HO\*) and peroxyl (R–O\*) radicals from the homolytic cleavage of PAA [2,8,27]. Therefore, this study investigates, for the first time, the potential benefits of coupling UV-C photolysis with PFA (UV-C/PFA) to enhance the reactivity of pharmaceuticals in phosphate buffer solution and wastewater effluent, using kinetic and removal rate approaches. As described previously [22], PFA exists as an equilibrium mixture of formic acid, water, and two oxidants: PFA and H₂O₂. Thus, UV-C/PFA generates a background AOP (UV-C/H₂O₂), whose contribution to the removal of pharmaceuticals by UV-C/PFA is investigated alongside a comparison of the two processes. In this study, the kinetic and removal rate approaches were investigated

for six pharmaceuticals: lidocaine (LID), furosemide (FUR), diclofenac (DCF), acetaminophen (ACT), sulfamethoxazole (SMX) and carbamazepine (CBZ), chosen due to their consistent detection in Parisian wastewater discharges [24], and their previously studied reactivity with PFA [46]. Except for LID and FUR, the other four pharmaceuticals exhibited very low reactivity with PFA.

The first necessary step involved evaluating the photolysis of PFA under UV-C irradiation, including its light absorption and quantum yield, and comparing it with PAA, as UV-C/PAA serves as the reference for comparing the two peracid-based AOPs. In addition, the reactive species responsible for pharmaceutical degradation under UV-C/PFA were qualitatively and systematically evaluated. This study also investigated the influence of initial pH on the reaction kinetics and degradation mechanisms of two pharmaceuticals: one PFA-reactive compound (lidocaine) and one less reactive compound (diclofenac).

To the best of the authors' knowledge, this study is the first to highlight the benefits of UV-C/PFA for pharmaceutical oxidation and systematically investigate the photolysis behavior of PFA during wastewater treatment.

# 2. Materials and methods

## 2.1. Chemical reagents

All reagents were used in their as-received state without additional purification (≥ 98 % purity for most reagents except for H₂O₂ (50 %) and phosphoric acid (85 %)). Pharmaceutical stock solutions were prepared in methanol (0.01–0.13 M), and their working solutions were obtained by dilution in deionized water (Milli-Q, 18 MΩ·cm). The PAA solution (38–40 % w/w) was a ready-to-use solution from Sigma-Aldrich. On the other hand, the PFA solutions were prepared in the laboratory by reacting formic acid (99 % purity) with H₂O₂ (50 % purity), following the method established previously [46,48]. The concentration of the peracid stock solutions was meticulously and periodically checked and found to contain, by weight, 19.2 ± 1.2 % of PFA (+ 19.5 ± 1.4 % of H₂O₂ and ~ 26 % of formic acid) or 28.4 ± 2.7 % of PAA (+ 4.3 ± 0.4 % of H₂O₂), using the iodometric titration method [12,56]. Due to its instability, the PFA stock solution was stored at –20 °C after preparation, while the PAA stock solution was stored at 4 °C. For specific details regarding the purity and suppliers of other reagents, refer to Text S1 in the *Supporting Information* (SI).

## 2.2. Collimated beam experiments

UV-C photolysis and AOP experiments (including UV-C/PFA, UV-C/PAA, and UV-C/H₂O₂) were conducted using a custom-made quasi-collimated beam apparatus equipped with three 16 W low-pressure mercury vapor lamps (Philips, Eindhoven, Netherlands, Figure S1, SI). These lamps emit at a principal monochromatic wavelength of 254 nm. The collimated beam apparatus was used to determine the photo-degradation kinetic rate constants of peracids and pharmaceuticals, as well as the removal rates of pharmaceuticals. The average UV-C intensity was determined with atrazine chemical actinometry (details described in SI, Text S2). The UV-C lamp was pre-heated for at least 30 min before each experiment to maintain stable light intensity, and continuous magnetic stirring was provided to ensure the reactor was thoroughly mixed at room temperature (~ 20 °C).

## 2.3. UV-C photolysis of PFA and PAA

PFA and PAA solutions were prepared at different concentrations in
deionized (DI) water without pH adjustment (at pH levels of 3.8 and 5.9, respectively, for PFA and PAA) and in a 10 mM phosphate buffer solution (PBS) at pH levels of 7.1, 7.7 and 9.2. A precise volume of peracid was injected individually into 200 mL of DI water or PBS under stirring in a crystallizer to achieve each initial concentration.

As a crucial step, the molar absorption coefficients at 254 nm ($ε_{254}$) of PFA ($1.3 \times 10^{-2}$ M) and PAA ($1.6 \times 10^{-2}$ M) in DI water (pH 3.8 and 5.9, respectively, for PFA and PAA) and PBS (pH 9.2) were determined by using the Beer-Lambert law. This included measuring the total absorbance of peracid solutions containing $H_2O_2$, subtracting the $H_2O_2$ absorbance to isolate the specific absorbance of the peracid, and then dividing by the initial molar concentration of the peracid (Eq. (1)):

$$ε_{254} = \frac{A_{peracid} - A_{H_2O_2}}{[peracid]_0 \times L} \quad (1)$$

Where $ε_{254}$ is the molar absorption coefficient of PFA or PAA ($M^{-1}.cm^{-1}$), A is the absorbance of peracid solution or pure $H_2O_2$ solution at 254 nm obtained after monitoring a blank with DI water or PBS depending on the solvent used for the preparation of the peracid solution ($cm^{-1}$), $[peracid]_0$ is the initial molar concentration of PFA or PAA (M), and L is the optical path length (1 cm).

In addition, the UV-C photolysis rate constants for PFA (29 µM) and PAA (23 µM) were determined both in DI water (without pH adjustment) and in a 10 mM PBS at pH levels of 7.1 and 7.7 to mimic a range of environmental conditions encountered in wastewater treatment. To determine the photolysis rate constants (Eq. (2)) and the quantum yield (Eq. (3)) of peracids, direct UV-C photolysis was performed with tert-butyl alcohol (TBA, 10 mM) added at the beginning of the reaction as an HO• scavenger [8]. The solution was then placed under the collimated beam apparatus within five seconds after peracid addition. For these experiments, the fluence rate was $0.3 \text{ mW.cm}^{-2} (1.1 \times 10^{-7} \text{ Einstein.L}^{-1}.\text{s}^{-1})$, conversion detailed in Text S2). Samples were taken at various intervals to quantify the residual peracid concentration using the N, N'-diethyl-p-phenylenediamine (DPD) colorimetric method [46], as the iodometric titration method is not suitable for low peracid concentrations. The photolytic decomposition of peracids followed first-order kinetics (Eq. (2)):

$$\frac{d[peracid]}{dt} = - k_{peracid} \times [peracid] \quad (2)$$

Where $k_{peracid}$ is the observed first-order photolytic rate constant for PFA or PAA ($min^{-1}$), and $[peracid]$ is the mass concentration of PFA or PAA ($mg.L^{-1}$).

In addition, the quantum yields of both peracids were calculated based on Eq. (3) (assuming the peracid is the only significant absorber of light), as previously described in the literature [8]:

$$\Phi_{254} = \frac{(k_{UV-C/peracid+TBA} - k_{peracid} - k_{UV-C/H_2O_2}) \times A}{I_{254} \times (1 - 10^{-A \times L}) \times \varepsilon_{254}} \quad (3)$$

Where $Φ_{254}$ is the quantum yield at 254 nm (mol.Einstein$^{-1}$), $k_{UV-C/peracid+TBA}$, $k_{peracid}$, $k_{UV-C/H_2O_2}$ are the rate constants experimentally determined ($s^{-1}$) for the direct UV-C photolysis of the peracids (after quenching with TBA), peracid autodecomposition, and photodegradation of $H_2O_2$ contained in the peracid solution, respectively. $I_{254}$, referred to as fluence, is the incident light intensity (Einstein.L$^{-1}.s^{-1}$), and L is the light path in water (3 cm in this study). Note that here, $A = (ε_{PFA} \times PFA) + (ε_{PFA} \times PFA) + (ε_{H_2O_2} \times H_2O_2)$ ($cm^{-1}$). All experiments were performed at least in duplicate.

## 2.4. Photodegradation experiments

The PBS (10 mM, pH 7.0) containing each pharmaceutical (LID, FUR, DCF, ACT, SMX and CBZ at 1 µM) was prepared in a 100 mL reactor. Then, a precise volume of PFA working solution was added to the reactor under stirring to achieve an initial concentration of 500 µM (31 mg.L$^{-1}$).

Within five seconds after the PFA addition, the reactor was placed under the collimated beam apparatus to initiate the AOP experiments. To evaluate the contribution of the background $H_2O_2$ concentration present in the PFA solution, control experiments were conducted with UV-C/$H_2O_2$ by substituting the PFA solution with pure $H_2O_2$ solution ($892 \text{ µM} = 30 \text{ mg.L}^{-1}$, corresponding to the background $H_2O_2$ concentration in $500 \text{ µM} = 31 \text{ mg.L}^{-1}$ of PFA). In addition, control experiments were carried out with UV-C photolysis (without PFA or pure $H_2O_2$) to assess its contribution to the degradation of pharmaceuticals by different AOPs.

Furthermore, the impact of the PFA initial concentration was evaluated on the production of HO•, given that these radicals were generated via UV-C/PAA [8,75,76,2] and expected to be also generated via UV-C/PFA. The HO• production was indirectly measured through the degradation of a quenching probe compound, parachlorobenzoic acid (pCBA, 1 µM), in 10 mM PBS (pH 7.0) with four different PFA concentrations: 16, 161, 500 and 806 µM. Additionally, the contribution of $H_2O_2$ to the production of HO• within the UV-C/peracid system was assessed. The kinetic rate constants of pCBA were thus compared between three systems: UV-C/PFA (500 µM, 31 mg.L$^{-1}$), UV-C/PAA (408 µM, 31 mg.L$^{-1}$) and UV-C/$H_2O_2$ ($H_2O_2$ concentrations equivalent to that in peracid solution: 892 µM and 147 µM as present in the PFA and PAA solutions, respectively). To check the contribution of each process individually on the HO• production, control pCBA degradation experiments were conducted by UV-C photolysis alone or with each oxidant alone (PFA or PAA or $H_2O_2$ without UV-C photolysis).

Furthermore, the contribution of some potentially produced reactive species to the degradation of pharmaceuticals by UV-C/PFA (500 µM of PFA) was evaluated on three pharmaceuticals (LID, DCF, and ACT at 1 µM) through quenching experiments using TBA and MeOH at 200 mM each.

In addition, the effect of pH was also evaluated during the oxidation of LID and DCF by UV-C/PFA (1 µM of pharmaceutical and 500 µM of PFA). This evaluation involved adjusting the initial pH of the PBS to five levels: 2.0, 5.0, 7.0, 9.0, and 11.0.

For all experiments, the reaction time in the reactor was 30 min for UV-C photolysis, PFA, and $H_2O_2$. It varied between 3 and 10 min for AOP experiments (UV-C/PFA and UV-C/$H_2O_2$), except for CBZ, which consistently underwent a 30 min oxidation across all processes due to its lower reactivity. For these experiments, the fluence was $0.5 \pm 0.01 \text{ mW.cm}^{-2} (2.3 \pm 0.03 \times 10^{-7} \text{ Einstein.L}^{-1}.\text{s}^{-1})$. At each UV-C dose interval, 1.5 mL of the sample was withdrawn from the reactor, injected into an amber vial containing excess sodium thiosulfate ($[Na_2S_2O_3]_0/[PFA]_0 = 7$) to stop the reaction and stored at 4°C before being analyzed (part 2.7). All experiments were conducted at least in duplicate, and the average results are presented.

Pharmaceuticals and pCBA photodegradation followed pseudo-first-order kinetics, and the observed kinetic rate constants ($min^{-1}$) were calculated following Eq. (4).

$$\frac{d[compound]}{dt} = \ln \frac{[compound]}{[compound]_0} = - k_{obs} \times [compound] \quad (4)$$

Where $k_{obs}$ is the observed pseudo-first-order kinetic rate constant for pharmaceuticals or pCBA with each process (UV-C/PFA, UV-C/PAA, UV-C/$H_2O_2$, UV-C photolysis and each oxidant alone for pCBA), [compound] and [compound]$_0$ are the molar concentrations of pharmaceuticals or pCBA at time t and initial time, respectively.

## 2.5. Identification of different reactive species produced by UV-C/PFA

The identification of reactive species produced by UV-C/PFA and potentially involved in the photodegradation of pharmaceuticals was assessed in PBS (pH 7.0) via kinetic experiments using Electron Para-magnetic Resonance (EPR) and mass spectrometry analysis. In this first experimental setup, 5,5-dimethyl-1-pyrroline-N-oxide (DMPO) and
2,2,6,6-tetramethyl-4-piperidol (TMP) were used as free radicals and singlet oxygen (¹O₂) trapping agents, respectively. Specifically, DMPO (2 and 50 mM) and TMP (50 mM) were oxidized by UV-C/PFA (1 mM, 62 mg.L⁻¹) and UV-C/H₂O₂ (1.8 mM, 61 mg.L⁻¹, corresponding to the background H₂O₂ concentration in the PFA solution). Moreover, specific DMPO experiments employed TBA (5 and 50 mM) and MeOH (100 mM) as quenching agents to distinguish HO• and some other organic radicals. In the second experimental setup, HO• production was indirectly quantified by monitoring the degradation of pCBA (1 µM) under UV-C/ PFA (500 µM PFA), while ¹O₂ generation was assessed using 9,10-diphenylanthracene (DPA, 0.1 mM) and furfuryl alcohol (FFA, 1.2 mM) under UV-C/PFA (1.0 mM PFA). The UV-C fluence was 0.5 mW.cm⁻² for TMP, DPA, and FFA experiments and 1.2 mW.cm⁻² for DMPO experiments, as the latter were conducted with a specific apparatus that did not allow for controlling the fluence. Control experiments were conducted with PFA and/or H₂O₂.

## 2.6. Removal of pharmaceuticals by UV-C/PFA in wastewater effluent

The photodegradation of the six pharmaceuticals was assessed in treated biological wastewater effluent collected from the Seine Amont WWTP located in Valenton, France. Samples were collected directly before the discharge channel and filtered in the laboratory through a 0.45 µm glass fiber filter (Whatman). The WWTP processes were previously described [48]. The wastewater quality parameters were as follows (mean ± standard deviation over three samples): pH 7.9 ± 0.1, total suspended solids 5.5 ± 2.8 mg.L⁻¹, UV absorbance at 254 nm (UV₂₅₄) 1.1 ± 0.6, nitrite 0.6 ± 0.1 mgN.L⁻¹, ammonium 1.6 ± 1.0 mg/L, phosphate 3.1 ± 0.9 mgP.L⁻¹, dissolved organic carbon 8.1 ± 0.7 mgC.L⁻¹, and temperature 22.9 ± 0.6 °C. The six pharmaceuticals (LID, FUR, DCF, ACT, SMX, and CBZ) were spiked into treated wastewater effluent at an individual concentration of 1 µg.L⁻¹ in addition to their initial concentration. The effluent was then subjected to batch oxidation experiments without any pH adjustment. A precise volume of PFA stock solution was added to achieve an initial concentration of 161 µM (10 mg.L⁻¹) under constant stirring, followed by immediate UV-C irradiation within five seconds. To determine the pharmaceutical removal rate, after 60 min, excess Na₂S₂O₃ ([Na₂S₂O₃]₀/[PFA]₀ = 7) was added to stop the reaction. Samples were then withdrawn and stored at 4 °C before being extracted and analyzed, as described in part 2.7. All experiments were conducted in triplicate to ensure accuracy.

## 2.7. Analytical methods

A high-performance liquid chromatography equipped with a diode array detector (HPLC-DAD, SPD-M20A, Shimadzu) was used to analyze pharmaceuticals in DI water and determine the kinetic rate constants. The analytical parameters, including stationary and mobile phases, elution gradient, and pharmaceutical-specific operating wavelengths with maximum UV absorbance, were previously documented [46]. For pCBA, the analytical method for pharmaceuticals was applied, using a maximum UV absorbance wavelength of 234 nm. All details concerning the analysis of atrazine, used to determine UV-C fluence, are provided in SI, Text S2.

The analysis of the six pharmaceuticals spiked into wastewater effluent was conducted using ultrahigh-performance liquid chromatography (UPLC) coupled with tandem mass spectrometry (MS/MS) employing a triple quadrupole detector (Acquity-TQD, Waters) equipped with an electrospray ionization (ESI) source. Samples were subjected to extraction using an automated solid-phase extraction (SPE) system (Dionex Autotrace 280, Thermo Scientific) using homemade multilayer SPE cartridges, as previously described [30,48]. Specific analytical parameters, including stationary and mobile phases, internal standards, elution gradient, ionization mode, and MS/MS acquisition parameters, were previously described [46].

A UPLC system, coupled to a high-resolution ion-mobility time-of-

flight mass spectrometry (Vion IMS-QTof, Waters) equipped with an ESI source, was used to analyze pharmaceutical transformation products (TPs), DPA, FFA, and their respective TPs (DPA endoperoxide and 6-hydroxy(2 H)pyran-3(6 H)one). The analysis was performed in both negative and positive ionization modes, and mass spectra were acquired in HDMS<sup>E</sup> (data independent analysis) mode to obtain low and high collision energy fragments. The instrument and methods were previously described [48].

The EPR spectra of reactive species were recorded in situ using a Bruker-Elexsys E500 spectrometer (Bruker, Germany) with an X Band frequency in continuous wave at room temperature. The settings for the EPR spectrometer were as follows: a center field of 3510 G, a sweep width of 100 G, a microwave frequency of 9.854 GHz, a microwave power of 20 mW, a receiver gain of 50 dB, a sweep time of 21 s, 1 G field modulation amplitude, and 100 KHz field modulation.

The absorption spectrum and the absorbance of all compounds were measured using a UV-Vis spectrometer (VWR Spectrophotometer UV-6300PC), and the pH of the solution was measured using a pH meter (VWR WTW-pH 3110 SenTix® 41).

# 3. Results and discussion

## 3.1. UV-C photolysis of peracids

To elucidate the degradation mechanisms of pharmaceuticals via UV-C/PFA, the photolysis of PFA at 254 nm was first investigated across three different pH levels. In experiments conducted in DI water (without pH regulation), the pH was 3.8, whereas in experiments conducted in PBS (10 mM), the pH was regulated to 7.1 and 7.7, and in these three conditions, the final pH decreased by less than 0.5 units. The UV-C photolysis of PFA followed first-order kinetics (Fig. 1a), consistent with its autodecomposition behavior [46]. In DI water at pH 3.8, the UV-C photolysis rate constant of PFA was 1.2 ± 0.03 × 10⁻¹ min⁻¹, while in PBS, it increased up to 6.3 ± 1.2 × 10⁻¹ and 5.8 ± 0.5 × 10⁻¹ min⁻¹, respectively at pH 7.1 and 7.7. Compared to PFA autodecomposition, UV-C photolysis increased PFA decay by a factor of 9–14 at the three tested pH levels (k<sub>autodecomposition</sub> = 8.7 ± 0.1 × 10⁻³ min⁻¹ in DI water at pH 3.8, 6.7 ± 0.4 × 10⁻² and 6.5 ± 0.5 × 10⁻² min⁻¹ in PBS respectively at pH 7.1 and 7.7, [46]). Therefore, the increased PFA decay under UV-C photolysis could be attributed to the co-existence of three mechanisms: autodecomposition, direct UV-C photolysis, and decomposition by the produced reactive species. Based on the reaction chain proposed for UV-C/PAA [8], the PFA photolysis could occur through a chain of radical reactions expressed by Eq. 5.

$$HC(=O)O_2H \xrightarrow{h\nu} HC(=O)O^* + HO^* \quad (a)$$

$$H_2O_2 \xrightarrow{h\nu} 2HO^* \quad (b)$$

$$HC(=O)O^* \rightarrow H^* + CO_2 \quad (c)$$

$$HC(=O)O_2H + HO^* \rightarrow HC(=O)OH + HOO^* \quad (d)$$

$$HC(=O)O_2H + HO^* \rightarrow HC(=O)O_2^* + H_2O \quad (e)$$

$$HC(=O)O_2H + HO^* \rightarrow HC(=O)O^* + O_2 + H_2O \quad (f)$$

$$H^* + O_2 \rightarrow HO_2^* \quad (g)$$

$$HC(=O)O_2H + HC(=O)O^* \rightarrow HC(=O)O_2^* + HC(=O)OH \quad (h)$$

$$HO^* + HO_2^* \rightarrow H_2O + O_2 \quad (i)$$

$$2HC(=O)O_2H \rightarrow 2HC(=O)OH + O_2^* \quad (j)$$

Eq. 5: Chemical reaction chain for the photolysis of PFA (adapted from
<figure><img src="image_3.png" /><figcaption>**Fig. 1.** Photolytic decomposition of PFA (a) and PAA (b) in 10 mM phosphate buffer solution (PBS) at various initial pH levels. The runs at pH 3.8 and 5.9 were carried out in DI water without pH regulation. [PFA or PAA]<sub>0</sub> = 1.8 ± 0.3 mg.L<sup>-1</sup> (equivalent to 29 µM of PFA and 23 µM of PAA).</figcaption></figure>

Cai et al., [9]

Initially, UV-C light induces the homolytic cleavage of the O–O bond in the PFA molecule, producing a formyloxy radical (HC(=O)O•) and HO• (a). In addition, the UV-C photolysis of H₂O₂ is an important source of HO• (b). The HC(=O)O• can decompose into a hydrogen radical (H•) and carbon dioxide (CO₂, c), while HO• can attack PFA to generate formic acid (HC(=O)OH) and hydroperoxyl radical (HOO•, d) or per-formyloxy radical (HC(=O)O•₂) and H₂O (e) or formyl radical (HC(=O)•), oxygen (O₂), and H₂O (f). Moreover, the generated H• (c) can react with O₂ (f) to form a hydroperoxyl radical (HO•₂, g), while HC(=O)O• can react with PFA, leading to the formation of HC(=O)O•₂ and HC(=O)OH (h). Additionally, HO• can react with HO•₂ (i), producing H₂O and ¹O₂. The formation of ¹O₂ also occurs via the autodecomposition of PFA (Eq. 5j and part 3.2) through an ionic mechanism [59,61].

Furthermore, the UV-C photolysis of PAA at 254 nm was also conducted for comparison purposes. Under acidic conditions (Fig. 1b, pH 5.9 resulting from the addition of PAA in DI water), the UV-C photolysis of PAA primarily relied on its slow autodecomposition, with a rate constant value ($k_{photolysis}$ = $3.4 \times 10^{-3}$ min$^{-1}$) closely aligned with the autodecomposition rate constant value previously reported at the same pH ($k_{autodecomposition}$ = $3.6 \times 10^{-3}$ min$^{-1}$, [46]). However, under neutral and alkaline conditions, besides the autodecomposition, both UV-C photolysis and decomposition by reactive species contributed to PAA decay, with $k_{photolysis}$ values of $1.6 \times 10^{-2}$ min$^{-1}$ (pH 7.1) and $2.0 \pm 0.06 \times 10^{-2}$ min$^{-1}$ (pH 7.7), approximately five times higher than $k_{autodecomposition}$ at pH 7.7 ($3.1 \times 10^{-3}$ min$^{-1}$, [46]).

Globally, the $k_{photolysis}$ for both PFA and PAA exhibited a similar trend to $k_{autodecomposition}$, with the decay occurring approximately five to six times faster under neutral and alkaline than under acidic conditions (Fig. 1). Thus, the higher $k_{photolysis}$ of PFA observed in neutral and

alkaline conditions could be attributed to its increased molar UV absorption coefficient at $\lambda$= 254 nm in alkaline ($\epsilon_{PFA}$ = $12.5 \pm 0.6$ M$^{-1}$. cm$^{-1}$ at pH 9.2) compared to acidic conditions ($\epsilon_{PFA}$ = $8.8 \pm 0.3$ M$^{-1}$. cm$^{-1}$ at pH 3.8). At pH > pKa = 7.1, $\epsilon_{PFA}$ should stabilize due to the predominance of its deprotonated form (PFA⁻) [46]. Nevertheless, the $\epsilon_{PAA}$ in acidic ($8.3 \pm 0.7$ M$^{-1}$.cm$^{-1}$ at pH 5.9) and alkaline ($7.7 \pm 0.5$ M$^{-1}$.cm$^{-1}$ at pH 9.2) conditions were quite similar, with the former closely matching the literature value ($8.0 \times 10^{-1}$.cm$^{-1}$, pH 5.1), whereas the last being notably lower than the literature value ($41.6 \times 10^{-1}$.cm$^{-1}$, pH 9.7, [9]), despite the higher $k_{photolysis}$ in alkaline conditions (Fig. 1). Moreover, the $\epsilon_{H2O2}$ at $\lambda$ = 254 nm was determined to be $19.0 \times 10^{-1}$.cm$^{-1}$, in accordance with the literature ($18.7 \times 10^{-1}$.cm$^{-1}$, [5]).

In addition, PFA was more susceptible to UV-C photolysis than PAA despite having nearly the same activation energy for decomposition: $95.4$ kJ.mol$^{-1}$ for PFA [61] and $92.1$ kJ.mol$^{-1}$ for PAA [72]. This higher $k_{photolysis}$ of PFA can be attributed to the elevated background concentration of H₂O₂ in the PFA solution ($1.8$ µM of H₂O₂ per µM of PFA solution vs. $0.4$ µM of H₂O₂ per µM of PAA solution), which could lead to greater production of HO• (part 3.3.1), subsequently accelerating the PFA decay (Eq. 5). Furthermore, this phenomenon is also corroborated by the higher apparent quantum yield ($\Phi$) of PFA, which is 0.9 mol. $Einstein^{-1}$ at pH 3.8 and $2.5$ mol.Einstein$^{-1}$ at pH 9.2, higher than the $\Phi$ of PAA ($0.7$ mol.Einstein$^{-1}$ at pH 5.9). The $\Phi_{PAA}$ is lower than the previously reported values of $0.88$ [76] and $1.20$ mol.Einstein$^{-1}$ [8], likely due to the lower fluence rate employed in the present study. Furthermore, the addition of TBA partially inhibited the UV-C photolysis of PFA and PAA, confirming peracid scavenging by HO•, with a more pronounced inhibitory effect observed under acidic conditions, as already reported for UV-C/PAA [8]. The reduced inhibitory effect of peracid photolysis under alkaline conditions after TBA addition is likely due to
the additional scavenging action of hydroxide ions (OH⁻) on the generated radicals, such as HO•, a mechanism previously documented [18]. Assuming that TBA scavenges only HO• and that HO• are primarily responsible for the indirect photolysis of peracids, the reaction kinetics between peracids and HO• were investigated (Text S3). The $k_{PFA/HO•}$ were $2.0 \times 10^9$ and $3.1 \times 10^9$ M⁻¹·s⁻¹, respectively, at pH 3.8 and 9.2, higher than the $k_{PAA/HO•}$ ($2.9 \times 10^8$ M⁻¹·s⁻¹ at pH 5.9). This $k_{PAA/HO•}$ was also lower than the value ($9.3 \times 10^8$ M⁻¹·s⁻¹ at pH 6.1) reported by Cai et al., [9], likely due to the lower fluence rate employed in the present study. Therefore, the $k_{PFA/HO•}$ are 1–2 orders of magnitude higher than $k_{PAA/HO•}$ and $k_{H2O2/HO•}$ ($2.7 \times 10^7$ M⁻¹·s⁻¹ as mentioned in Buxton et al., [7] without specifying pH, which was measured to be 5.2 in the present study), indicating that HO• exhibits a stronger scavenging effect on PFA. To the best of our knowledge, this is the first study to report the $\Phi_{PFA}$ and $k_{PFA/HO•}$. These data and the reactions mentioned earlier (Eq. 5) highlight the complexity of radical mechanisms involved in the activation of PFA by UV-C photolysis, generating various reactive species (HO•, R–O•, and $^1O_2$) that may exhibit different reactivities. This underlines the necessity for experimental validation using other techniques like EPR to confirm the presence of some reactive species (part 3.2).

## 3.2. Identification of reactive species by electron paramagnetic resonance (EPR)

Several techniques enable the identification and quantification of reactive species in AOPs. They include physical (e.g., spectral measurement of chemiluminescence using a red-sensitive photomultiplier or a liquid nitrogen-cooled germanium diode detector), and chemical techniques using different probe compounds, 9-[2-(3-carboxy-9,10-diphenyl)anthryl]-6-hydroxy-3H-xanthen-3-ones for fluorescence detection [23,62], DPA [43,44], FFA [26], pCBA for spectroscopic and spectrophotometric detection, 2,2,6,6-tetramethylpiperidin and its derivatives (e.g., 2,2,6,6-tetramethyl-4-piperidinol and 2,2,6,6-tetramethyl-4-piperidone) for EPR detection. In the present study, in situ EPR

measurement using spin trap compounds was chosen for two main reasons: first, it enables the detection of the two main hypothesized reactive species (HO• and $^1O_2$), and second, to the best of our knowledge, all studies focusing on the detection of reactive species in peracid systems, specifically peracetic acid, have employed this technique [2,18,33,36,37,57,66,68,70,73,74]. In this regard, free radicals and $^1O_2$ produced by PFA and UV-C/PFA were monitored using two probe compounds: DMPO and TMP, respectively. Additionally, to distinguish the free radicals, based on results on UV-C/PAA [18,32,8], TBA was used to selectively quench HO•, while MeOH was used to quench both HO• and some R–O• [8,18,32], although no relevant data exists regarding the capability of MeOH to quench the R–O• produced by UV-C/PFA. In addition, the production of $^1O_2$ and HO• was also confirmed by spectroscopic detection (parts 3.2 for $^1O_2$ and 3.3.1 for HO•).

After 5 min of reaction with DMPO, no radical signal was detected with PFA alone (Fig. 2a). However, characteristic peaks corresponding to HO• (DMPO–HO•, with hyperfine splitting constants of $a_N = a_H = 14.9$ G, a four-fold characteristic peak with intensities in the ratio 1:2:2:1) were observed with UV-C/PFA and UV-C/$^1H_2O_2$. For the same $^1H_2O_2$ concentration ($1.8$ mM = $61$ mg·L⁻¹, equivalent to the background concentration in PFA solution), the peak intensities were higher under UV-C/$^1H_2O_2$ than UV-C/PFA, implying that the former produced more HO•. This difference is likely due to the scavenging effect of formate ions ($1$ pKa = $3.75$) present in the PFA solution, as the experiments were conducted in PBS at pH $7.1 \pm 0.1$. This finding is consistent with previous observations for UV-C/PAA and UV-C/$^1H_2O_2$, where the former produces fewer HO• despite the higher $\Phi_{PAA}$ ($0.9$ mol·Einstein⁻¹) than $\Phi_{H2O2}$ ($0.5$ mol·Einstein⁻¹), due to the scavenging effect of acetate ions and R–O• in UV-C/PAA [76]. Furthermore, the addition of TBA ($5$ mM) in both systems (UV-C/PFA and UV-C/$^1H_2O_2$) had the same effect: nearly total suppression of the HO• signal (referring to the peak-to-peak amplitude after adjusting the baseline to zero, Fig. 2a).

In addition, the intensity of HO• increased over time under UV-C/PFA (Fig. 2b), reaching maximum intensity at $6$ min, then decreased from $6$ to $8$ min, and disappeared entirely at $10$ min. The decrease in

<figure><img src="image_4.png" /><figcaption>**Fig. 2.** Spin trapping EPR spectra of DMPO before and after oxidation by PFA, UV-C/PFA, UV-C/PFA + TBA, UV-C/$^1H_2O_2$ and UV-C/$^1H_2O_2$ + TBA (a), UV-C/PFA at different reaction times (b), and UV-C/PFA, UV-C/PFA + TBA, and UV-C/PFA + MeOH (c), with points corresponding to HO• and triangles to other organic radical signals (c). Conditions: [DMPO] = $2$ mM for a and b, $50$ mM for c, $[PFA]_0 = 1$ mM, $[^1H_2O_2]_0 = 1.8$ mM, $[TBA]_0 = 5$ mM for a and b, $50$ mM for c, $[MeOH]_0 = 100$ mM for c, initial pH $7.1 \pm 0.1$. The intensity refers to the peak-to-peak amplitude after adjusting the baseline to zero and is in an arbitrary unit.</figcaption></figure>
HO• production over time likely results from the increased PFA decomposition and HO• consumption through other radical reactions (Eq. 5). The evolution of HO• production under UV-C/PFA can affect the removal of pharmaceuticals reactive with HO•. Furthermore, additional radical peaks were observed under UV-C/PFA when the initial concentration of DMPO was increased to 50 mM (Fig. 2c, peaks marked by triangles). These peaks exhibited the following hyperfine splitting constants: aN = 15.4 G and aH = 23.0 G, which could correspond to the primary R–O•, specifically HCOO• and HCOOO• (Fig. 2c). The proposition of HCOO• and HCOOO• as the primary formed R–O• is based on their suggested presence in PFA-based AOPs [19] and the reactive radical structures previously suggested for UV-C/PAA [8,76]. Interestingly, the addition of TBA (50 mM) or MeOH (100 mM) slightly reduced the intensity of HO•and did not affect R–O• (Fig. 2c). This suggests that while MeOH is commonly used as a scavenger for several types of reactive radicals, including HO•, it is not suitable to quench all R–O• generated by UV-C/PFA. Some studies mentioned the use of 2,4-hexadiene instead of MeOH in PAA-based AOPs to achieve broader quenching results [32,76]. Further experiments are needed to evaluate the quenching impact of MeOH through UV-C/PFA (part 3.3.3). Therefore, the limited effect of TBA and MeOH on HO•, resulting in only a slight reduction of the HO• signal, is attributed to the high concentration of DMPO (50 mM), which scavenges HO• at a rate (3.4–4.3 × 10⁹ M⁻¹.s⁻¹, [1,20]) that surpasses the quenching efficiency of TBA (7.6 × 10⁸ M⁻¹.s⁻¹, [18]) and MeOH (9.7 × 10⁸ M⁻¹.s⁻¹, [18]). Furthermore, the generation of HO• via UV-C/PFA was also confirmed by the observed oxidation of pCBA (Fig. 3).

Additionally, the ¹O₂ signal (a third-fold characteristic peak with intensities in the ratio 1:1:1) corresponding to the 4-hydroxy-2,2,6,6-tetramethylpiperidin-1-oxyl (TEMPO), an oxidation product generated by the reaction between TMP and ¹O₂ [14,45], was detected in both PFA and UV-C/PFA systems (Figure S2) with the intensity increasing over time (Figure S3), while no signal was observed in the TMP solution. A control experiment confirmed that phosphate ions did not affect the formation of ¹O₂, as evidenced by the intensity comparison between

experiments in DI water and PBS under identical conditions of PFA concentration and reaction time (Figure S3). These results suggest that ¹O₂ was generated through PFA autodecomposition (Eq. 5j), consistent with prior findings on PAA [18]. Moreover, UV-C/PFA showed a significantly higher peak intensity, indicating enhanced ¹O₂ formation via the reaction between HO• and HO₂• (Eq. 5i). ¹O₂ formation was further investigated by oxidizing DPA and FFA under PFA and UV-C/PFA. This approach was explored due to the potential side reactions between TMP and radicals, such as HO•, which may reduce the EPR signal intensity [54]. The detection of DPA endoperoxide and 6-hydroxy(2 H)pyran-3(6 H)one (Figure S4a and b), specific TPs of DPA and FFA, formed via oxidation by ¹O₂ [26,43,44], confirmed the generation of ¹O₂. Their peak intensities increased over time, with higher levels observed in UV-C/PFA (Figure S4c and d), supporting the EPR results. Although FFA and DPA confirmed ¹O₂ formation, their suitability as probes in PFA-based AOPs is limited. While FFA reacts faster with HO• (1.5 × 10¹⁰ M⁻¹.s⁻¹, [77]) than with ¹O₂ (1.0 × 10⁸ M⁻¹.s⁻¹, [3]), compromising its selectivity, DPA has poor water solubility and is photosensitive [8,78,71]. Furthermore, anthracene-derived endoperoxides can photodecompose back to anthracene and other products [55], restricting their applicability in photo-AOPs. Given the limited selectivity and specificity of existing probes in peracid-based AOPs, improved and more specific ¹O₂ probes are needed.

However, the ¹O₂ produced through PFA autodecomposition can explain its selective reactivity with specific functional groups, such as reduced sulfur, aniline, deprotonated tertiary amine, and furan rings, as demonstrated previously [46]. ¹O₂ is indeed known for its high reactivity towards compounds containing electron-rich sites, leading to electrophilic addition, single-electron transfer, or oxygen addition to heteroatoms, specifically nitrogen or sulfur [17,41,54]. In addition, ¹O₂ also preferentially attacks the ortho and para positions of phenol rings due to the higher electron density at these sites, resulting from the ortho-para directing nature of the –OH group. This is in accordance with the increased reactivity of PFA with catechol and hydroquinone compared to resorcinol, as previously reported [46].

<figure><img src="image_5.png" /><figcaption>**Fig. 3.** Effect of PFA concentration on the rate constant of pCBA degradation by UV-C/PFA. Experimental conditions: [pCBA]₀ = 1 µM in phosphate buffer solution (10 mM, pH 7.0), [PFA]₀ = 16–806 µM, [H₂O₂]₀ = 29–1 439 µM, 30 min of reaction, UV-C fluence = 0.5 ± 0.01 mW.cm⁻² = 2.3 ± 0.03 × 10⁻⁷ Einstein. L⁻¹.s⁻¹.</figcaption></figure>

### 3.3. Degradation of organic compounds by UV-C/PFA

#### 3.3.1. Quantification of HO• produced through UV-C/PFA and UV-C/PAA

To determine the optimal PFA concentration for subsequent experiments, HO• production through UV-C/PFA was investigated. HO• are known for their non-selectivity and high oxidation efficiency toward numerous micropollutants [69], owing to their high redox potential of 2.8 V/ESH [49] and can be easily measured with probe compounds. Thus, HO• were indirectly quantified using pCBA as the probe compound ($k_{pCBA/HO•} = 5.0 \times 10^9 M^{-1}.s^{-1}$, [58]) because previous studies have demonstrated that pCBA cannot be oxidized by the organic radicals produced specifically by PAA-based AOPs [15,65]. Due to the high reactivity of pCBA with HO•, its rapid degradation, expressed by a first-order rate constant (min⁻¹), serves as an indicator of HO• abundance. Prior to quantifying HO• generated by UV-C/PFA and UV-C/PAA, control experiments were performed to determine whether each process independently produced HO•. To this end, the degradation of pCBA (1 µM) was initially evaluated under UV-C photolysis and in the presence of individual oxidants (PFA or PAA or H₂O₂ at 31 mg.L⁻¹). The results showed that pCBA degradation by these processes was very low (Figure S5), near zero, with rate constants of 3.8 ± 0.7 × 10⁻³, 9.9 ± 4.2 × 10⁻⁴, 9.4 ± 3.7 × 10⁻⁴ and 6.8 ± 0.4 × 10⁻⁴ min⁻¹ for UV-C photolysis, PFA, PAA and H₂O₂, respectively. This finding, in accordance with EPR results, confirms that in the present experimental conditions, none of these processes could individually produce HO• and contradicts the assumption made in the literature that peracid disinfection relies on a radical mechanism involving reactive species such as HO• [22,31,38,39,60]. Additionally, the lack of pCBA degradation by PFA suggests that the ¹O₂ generated through PFA autodecomposition does not react with pCBA. Then, pCBA (1 µM) was oxidized by
UV-C/PFA using four PFA concentrations: 16, 161, 500, and 806 µM (equivalent to 1, 10, 31, and 50 mg.L⁻¹) under a UV-C fluence of 0.5 ± 0.01 mW.cm⁻² (2.3 ± 0.03 × 10⁻⁷ Einstein.L⁻¹.s⁻¹). These PFA concentrations contained approximately equivalent mass concentrations of H₂O₂ (29, 288, 892, and 1439 µM, equivalent to 1, 10, 30, and 49 mg. L⁻¹). Fig. 3 demonstrates that, under UV-C/PFA, the pCBA degradation increased with PFA concentration from 16 to 500 µM, after which it stabilized up to 806 µM. This increase in HO• production can be linked to the simultaneous increase of both PFA and background H₂O₂ concentration. However, the stabilization of HO• production at higher PFA concentrations (> 500 µM) can be attributed to scavenging effects resulting from potential side reactions between PFA or background H₂O₂ and the generated reactive species, including HO•. This phenomenon is comparable to that occurring in the UV-C/H₂O₂ and UV-C/PAA processes (considering its UV-C/H₂O₂ background AOP), where high oxidant concentrations (e.g., > 0.59 mM of H₂O₂ or 0.10 mM of PAA) can act as active scavengers. Indeed, H₂O₂ and PAA are renowned for their strong reactivity with HO•, with rate constants of kH₂O₂/HO• of 2.7 × 10⁷ M⁻¹.s⁻¹ at pH 5.2 [7] and kPAA/HO• of 2.9 × 10⁸ M⁻¹.s⁻¹ at pH 5.9, often leading to the formation of less reactive radicals such as HO•² in the case of H₂O₂ [11,35]. As an increase in PFA concentration leads to an increase in H₂O₂ concentration (in the same ratio), it is worth considering the existence of an optimal H₂O₂ concentration in the UV-C/PFA process, as it has been observed in other AOPs such as UV-C/H₂O₂ [35,52] or ozone/H₂O₂ [13]. Thus, the excess H₂O₂ concentration might be photon flux dependent and could potentially lead to the consumption of the produced HO•, subsequently decreasing HO• yield. Therefore, 500 µM of PFA (containing 892 µM of H₂O₂) was chosen as the optimal concentration for HO• production for the subsequent experiments.

Additionally, this experiment aimed to determine the contribution of H₂O₂ to HO• production through peracid-based AOPs. The production of HO• was thus compared between the UV-C/peracid (31 mg.L⁻¹ of peracids = 500 µM of PFA and 408 µM of PAA) and UV-C/H₂O₂ systems (Fig. 4) at the H₂O₂ concentrations equivalent to that present in peracid solutions (i.e., 892 µM in PFA and 147 µM in PAA, respectively). The results demonstrated significantly higher pCBA degradation under UV-

<figure><img src="image_6.png" /><figcaption>**Fig. 4.** Contribution of H₂O₂ on HO• production in UV-C/peracid system (through UV-C/H₂O₂ at equivalent concentrations of H₂O₂). [pCBA]₀ = 1 µM in phosphate buffer solution (10 mM, pH 7.0), [PFA]₀ = 500 µM ([H₂O₂]₀ = 892 µM), [PAA]₀ = 408 µM ([H₂O₂]₀ = 147 µM), 30 min of reaction, UV-C fluence = 0.5 ± 0.01 mW.cm⁻² = 2.3 ± 0.03 × 10⁻⁷ Einstein.L⁻¹.s⁻¹.</figcaption></figure>

C/PFA than UV-C/PAA, and showed that HO• production under UV-C/H₂O₂ was comparable to that observed with UV-C/PFA and UV-C/PAA (Fig. 4).

The ratio between the kinetic rate constants for pCBA degradation under UV-C/PFA and UV-C/PAA is 4, which closely matches the ratio between the background concentrations of H₂O₂ in peracids, equivalent to 6. This suggests that HO• under UV-C/PFA or UV-C/PAA primarily originate from the photolysis of background H₂O₂ in the peracid solution, contrary to previous literature, which indicated that HO• production under UV-C/PAA mainly arise from PAA photolysis [75]. If PAA or PFA can indeed produce HO•, these radicals are present in negligible amounts compared to the quantities produced by the background H₂O₂ concentration. As far as we know, this is the first time the contribution of H₂O₂ in UV-C/PFA or UV-C/PAA has been demonstrated and quantified in terms of HO• production. Therefore, HO• was produced more under UV-C/PFA than UV-C/PAA, as the H₂O₂ background concentration is higher in the PFA (~ 20 % w/w) than in the PAA (~ 4 % w/w) solution.

### 3.3.2. Kinetic rate constants of pharmaceuticals under UV-C/PFA in phosphate buffer solution

Six pharmaceuticals (LID, FUR, DCF, ACT, SMX and CBZ) at 1 µM were oxidized by UV-C/PFA (500 µM of PFA, UV-C fluence of 0.5 mW. cm⁻² = 2.3 ± 0.03 × 10⁻⁷ Einstein.L⁻¹.s⁻¹) in PBS (10 mM, pH 7.0). The plot of ln([pharmaceutical]/[pharmaceutical]₀) vs. time resulted in a straight line (data not shown), confirming that the photolytic degradation of pharmaceuticals by UV-C/PFA followed apparent first-order kinetics. Pharmaceutical solutions contained 0.03–0.05 % MeOH. Although previous studies [16] indicated that MeOH may influence HO• generation and thus affect AOP efficiency, the low concentration used in this study minimizes such effects. For instance, the kinetic rate constants for DCF degradation under UV-C/H₂O₂ closely align with literature values [29].

A control experiment with UV-C photolysis alone was conducted using the six pharmaceuticals and revealed kinetic rate constants (kUV-C) varying between 1.4 ± 0.3 × 10⁻⁴ min⁻¹ for CBZ and 2.9 ± 0.03 × 10⁻¹ min⁻¹ for DCF. Since UV-C/H₂O₂ is the primary source of HO• (part 3.3.1) in the UV-C/peracid system, a supplementary experiment was conducted by using pure H₂O₂ at the same concentration as that in the PFA solution (892 µM of H₂O₂ in 500 µM of PFA) to investigate its contribution to the degradation of pharmaceuticals. The kUV-C values of DCF, SMX, and FUR were close to their kUV-C/H₂O₂, indicating that direct UV-C photolysis played a significant role in their degradation compared to HO• produced from UV-C/H₂O₂ (kUV-C/H₂O₂/kUV-C ratio of 1.0, 1.1, and 1.3, respectively). The experiment also showed that CBZ degradation was due to HO• rather than direct UV-C photolysis (kUV-C/H₂O₂/kUV-C ratio of 30.3). Additionally, after 5 min of reaction under UV-C/PFA, DCF, SMX, FUR, and ACT exhibited degradation rates of ≥ 80.0 %, whereas their degradation was very low (~ 10.0 % after 30 min) with PFA alone. Under UV-C/PFA (5 min of reaction), LID was degraded at only 45 %, while CBZ displayed the lowest reactivity, with a degradation rate of < 4.0 %. Therefore, the UV-C/PFA kinetic rate constants for all six pharmaceuticals ranged from 4.8 ± 0.3 × 10⁻³ min⁻¹ for CBZ to 5.5 ± 0.1 × 10⁻¹ min⁻¹ for ACT (Table 1). Compared to the kPFA of these pharmaceuticals (Table 1, data obtained from Nabintu Kajoka et al., [46]), the kUV-C/PFA were approximately 22–171 times higher than kPFA, except for LID, whose kUV-C/PFA was only 1.2 times higher. Therefore, a synergistic effect between UV-C photolysis and PFA was observed for all pharmaceuticals except LID. This increased reactivity of pharmaceuticals with UV-C/PFA is likely attributed to the generation of various reactive species resulting from the activation of PFA. For all the investigated pharmaceuticals, the kUV-C/PFA/kUV-C/H₂O₂ ratios ranged from 1.2 (CBZ) to 14.9 (FUR), confirming that, apart from the HO• generated from the background H₂O₂, other reactive species (¹O₂ and R–O•) played a more significant role in the degradation of these pharmaceuticals. These additional reactive species reacted selectively with some pharmaceuticals, such as FUR, ACT, LID, and SMX, while exhibiting lower
<caption>**Table 1**</caption>

The pseudo-first-order rate constant ($k_{obs} \text{min}^{-1}$) of pharmaceuticals under PFA (data from [46]), UV-C/PFA, UV-C/H$_2$O$_2$, and UV-C alone (this work). Experimental conditions: [pharmaceutical]$_0$ = 1 µM, [PFA]$_0$ = 500 µM, [H$_2$O$_2$]$_0$ = 892 µM in UV-C/H$_2$O$_2$ experiments, [PBS] = 10 mM, pH 7.10, 20 ± 1 °C, UV-C fluence = 0.5 ± 0.01 mW.cm$^{-2}$ = 2.3 ± 0.03 × 10$^{-7}$ Einstein.L$^{-1}$.s$^{-1}$.

<table><thead><tr><th rowspan="2">Compounds</th><th rowspan="2">k<sub>PFA</sub></th><th rowspan="2">k<sub>UV-C/PFA</sub></th><th rowspan="2">k<sub>UV-C/H<sub>2</sub>O<sub>2</sub></sub></th><th rowspan="2">k<sub>UV-C</sub></th><th colspan="2">k<sub>UV-C/PFA</sub></th><th colspan="2">k<sub>UV-C/H<sub>2</sub>O<sub>2</sub></sub></th></tr><tr><th>k<sub>PFA</sub></th><th>k<sub>UV-C/H<sub>2</sub>O<sub>2</sub></sub></th><th>k<sub>UV-C</sub></th><th>k<sub>UV-C/H<sub>2</sub>O<sub>2</sub></sub></th></tr></thead><tbody><tr><td>LID</td><td>(9.3 ± 0.9) x10<sup>-2</sup></td><td>(1.1 ± 0.1) x10<sup>-1</sup></td><td>(2.6 ± 0.5) x10<sup>-2</sup></td><td>(1.3 ± 0.3) x10<sup>-4</sup></td><td>1.2</td><td>4.1</td><td>788.5</td><td>193.1</td></tr><tr><td>FUR</td><td>(1.2 ± 0.5) x10<sup>-2</sup></td><td>(3.8 ± 0.9) x10<sup>-1</sup></td><td>(2.5 ± 0.2) x10<sup>-2</sup></td><td>(2.0 ± 0.02) x10<sup>-2</sup></td><td>31.5</td><td>14.9</td><td>18.8</td><td>1.3</td></tr><tr><td>SMX</td><td>(4.0 ± 0.7) x10<sup>-3</sup></td><td>(3.3 ± 0.5) x10<sup>-1</sup></td><td>(1.1 ± 0.2) x10<sup>-1</sup></td><td>(1.2 ± 0.01) x10<sup>-1</sup></td><td>83.3</td><td>2.9</td><td>2.8</td><td>1.0</td></tr><tr><td>DCF</td><td>(4.0 ± 1.1) x10<sup>-3</sup></td><td>(4.8 ± 0.9) x10<sup>-1</sup></td><td>(3.3 ± 0.4) x10<sup>-1</sup></td><td>(2.9 ± 0.03) x10<sup>-1</sup></td><td>120.4</td><td>1.4</td><td>1.6</td><td>1.1</td></tr><tr><td>ACT</td><td>(3.2 ± 0.7) x10<sup>-3</sup></td><td>(5.5 ± 0.1) x10<sup>-1</sup></td><td>(9.4 ± 1.7) x10<sup>-2</sup></td><td>(1.6 ± 0.2) x10<sup>-3</sup></td><td>170.5</td><td>5.8</td><td>346.5</td><td>59.7</td></tr><tr><td>CBZ</td><td>(2.1 ± 0.3) x10<sup>-4</sup></td><td>(4.8 ± 0.3) x10<sup>-3</sup></td><td>(4.1 ± 0.2) x10<sup>-3</sup></td><td>(1.4 ± 0.3) x10<sup>-4</sup></td><td>22.5</td><td>1.2</td><td>35.3</td><td>30.3</td></tr></tbody></table>

reactivity towards CBZ and DCF, as evidenced by their low $k_{UV-C/PFA}/k_{UV-C/H_2O_2}$ ratio (Table 1). Since HO• are known for their non-selectivity, if they were the sole primary reactive species generated by UV-C/PFA, the $k_{obs}$ for all pharmaceuticals would likely be within the same order of magnitude. However, no literature is currently available on UV-C/PFA; still, Cai et al., [9] studied the reactivity of six pharmaceuticals, including DCF and CBZ, with UV-C/PAA and found kinetic rate constants of 1.4 ± 0.03 and 8.3 ± 0.3 × 10$^{-2}$ min$^{-1}$, respectively, which were higher than those observed for UV-C/PFA (4.76 ± 0.85 × 10$^{-1}$ min$^{-1}$ and 4.76 ± 0.32 × 10$^{-3}$ min$^{-1}$, respectively). This difference may be attributed to the significantly higher fluence rate used in the study by Cai et al., [9] (2.1 × 10$^{-6}$ Einstein.L$^{-1}$.s$^{-1}$) compared to that employed in the present study (3.4 ± 0.1 × 10$^{-7}$ Einstein.L$^{-1}$.s$^{-1}$), which likely resulted in higher production of reactive species and/or also higher UV-C photolysis rates. Another study reported a $k_{obs}$ of ~ 0.1 min$^{-1}$ for DCF with UV-C/PAA, using a high light intensity (2.1 mW.cm$^{-2}$ = 9.7 × 10$^{-7}$ Einstein.L$^{-1}$.s$^{-1}$) [75].

In conclusion, the reactivity of all pharmaceuticals was influenced by either HO• (based on $k_{UV-C/H_2O_2}/k_{UV-C}$ in this order: LID > ACT > FUR > CBZ > DCF, with the last two pharmaceuticals being less impacted) or additional reactive species (including $^1O_2$ and organic radicals, based on the $k_{UV-C/PFA}/k_{UV-C/H_2O_2}$ ratio: FUR > ACT > LID > SMX > DCF ~ CBZ). Therefore, the oxidation of pharmaceuticals by UV-C/PFA is a complex, molecule-dependent process that involves direct UV-C photolysis, oxidation by PFA, and, in some cases, oxidation by various reactive species, including HO•, R-O•, and $^1O_2$. This experiment highlights the need to investigate further the contribution of these reactive species to the degradation of pharmaceuticals (part 3.3.3).

### 3.3.3. Contribution of reactive species in the degradation of pharmaceuticals by UV-C/PFA

As observed previously (part 3.3.2), UV-C/PFA enhances the reactivity of pharmaceuticals due to the formation of various reactive species (HO•, R-O•, and $^1O_2$). Kinetic experiments using TBA and MeOH as quenching agents were conducted to quantify the contribution of these reactive species to pharmaceutical degradation. LID, DCF, and ACT were selected as the model compounds to include one PFA-reactive compound (LID) and two PFA non-reactive compounds (DCF and ACT), with ACT known for its potential susceptibility to R-O• (Table 1). The contribution of reactive species was estimated based on the kinetic rate constants (min$^{-1}$) obtained under UV-C/PFA+TBA and UV-C/PFA+MeOH, as described in Eqs. (6)–(9).

$$ \%UV-C_{\text{photolysis}} = \frac{k_{UV-C}}{k_{UV-C/PFA}} * 100 \quad (6) $$

$$ \%HO^{\bullet} = \frac{k_{UV-C/PFA} - k_{UV-C/PFA+TBA}}{k_{UV-C/PFA}} * 100 \quad (7) $$

$$ \%R-O^{\bullet} = \frac{k_{UV-C/PFA+TBA} - k_{UV-C/PFA+MeOH}}{k_{UV-C/PFA}} \quad (8) $$

$$ \%\text{otherreactivespecies} = 100 - \%\text{UV-C}_{\text{photolysis}} - \%\text{HO}^{\bullet} - \%\text{R-O}^{\bullet} \quad (9) $$

TBA and MeOH had varying impacts on the three pharmaceuticals (Fig. 5). TBA significantly inhibited the degradation of LID and DCF but had little impact on ACT. The stronger inhibitory effect of MeOH on LID and ACT compared to TBA supports the hypothesis that, in addition to HO•, other reactive species played a more significant role in their degradation under UV-C/PFA (Table S1). Accordingly, HO• accounted for 32.9 %, 30.6 %, and 8.9 % of the degradation of LID, DCF, and ACT, respectively (Table S1), while some R-O• (additional radicals quenched by MeOH) contributed 45.6 %, 10.6 %, and 52.5 % to their degradation. In the case of DCF (a photosensitive compound with direct UV-C photolysis contributing at 52.7 %), the results (Table S1) overall corroborate previous findings under UV-C/PAA, where 56.0 % of its degradation was ascribed to direct UV-C photolysis, 32.0 % to HO•, and 12.0 % to other organic radicals [75].

Overall, these results (Fig. 5 and Table S1) support the earlier observations (part 3.3.2), where the simultaneous contribution of HO• and R-O• was observed for LID, and that among the three pharmaceuticals, the higher contribution of R-O• was highlighted for ACT. The percentage attributed to other reactive species reflects the contribution of PFA and $^1O_2$ for PFA-reactive compounds (LID), as well as all other reactive species not quenched by MeOH or TBA. In addition, further research is needed to isolate the contribution of $^1O_2$ alone by using other quenching agents, such as sodium azide. However, the lower value of $k_{UV-C/PFA+MeOH}$ for LID, compared to $k_{PFA}$, is attributed to the decreased availability of PFA for oxidation due to its photolytic decomposition. As a result, it was not possible to quantify the PFA molecule contribution under UV-C/PFA. It is hypothesized, however, that the PFA contribution is less significant under UV-C/PFA than when used alone, primarily due to the high photodegradation of PFA. This consideration is important when applying this AOP in real WWTPs, where using these processes sequentially rather than simultaneously could better leverage PFA’s selective reactivity for certain compounds (e.g., LID).

### 3.3.4. Influence of the initial pH and identification of transformation products in phosphate buffer solution

The pH influence was evaluated during the oxidation of LID and DCF by UV-C/PFA (1 µM of pharmaceutical and 500 µM of PFA). For this, the initial pH of the PBS was regulated to five levels: 2.0, 5.0, 7.0, 9.0, and 11.0. In each experiment, the final pH decreased by less than 0.5 units. This broad pH range (2.0–11.0) was selected to assess the influence of pH on the reactivity of UV-C/PFA, considering that the autodecomposition behavior of PFA was previously described across this pH range [46]. Additionally, this pH range encompasses the typical wastewater pH in the Paris region (average pH 7.9 ± 0.1), used here to evaluate pharmaceutical removal under real conditions (Part 3.4).

During the oxidation of LID by UV-C/PFA, as previously observed with its degradation by PFA, the $k_{UV-C/PFA}$ increased with increasing pH (from 2.0 to 8.0, stabilizing up to pH 11.0) due to the predominant presence of the deprotonated forms of both LID and PFA [46]. The analogous pH-dependent behavior in the oxidation of LID by both PFA and UV-C/PFA is primarily due to the significant contribution of the PFA molecule and $^1O_2$ to its degradation under UV-C/PFA (part 3.3.3) compared to other reactive species present in the system. In addition, the
<figure><img src="image_7.png" /><figcaption>**Fig. 5.** Contribution of UV-C photolysis, HO•, and additional reactive species to the degradation kinetics of lidocaine, diclofenac, and acetaminophen under UV-C/PFA. Experimental conditions: [pharmaceutical]₀ = 1 µM, [PFA]₀ = 500 µM, [TBA] = [MeOH] = 200 mM, [PBS] = 10 mM, pH 7.1, 20 ± 1 °C, UV-C fluence = 0.5 ± 0.01 mW.cm⁻² = 2.3 ± 0.03 × 10⁻⁷ Einstein.L⁻¹.s⁻¹.</figcaption></figure>

reactivity of DCF under UV-C/PFA increased to a maximum at pH 7.0 and decreased in alkaline conditions (pH > 7.0). The low reactivity of DCF in both acidic and alkaline conditions may be attributed to the consumption of HO•, which plays a significant role in the degradation of DCF (30.6 % of the $k_{UV-C/PFA}$, part 3.3.3) through side reactions with both $H^+$ and $OH^-$ ions. This phenomenon was previously observed during the degradation of propiconazole, a triazole fungicide, by UV-C/H₂O₂ [28]. A similar behavior was noted in the degradation of DCF by UV/TiO₂, where the authors primarily attributed this effect to the surface charge of TiO₂ [64] rather than the structure or properties of DCF. The recurrence of this phenomenon in UV-C/PFA suggests the need for further investigation into the role of direct UV-C photolysis in relation to DCF properties. Additional insights into the role of pH in the degradation of LID and DCF are discussed in Text S4 and supported by Figure S6.

In addition, to gain a deeper understanding of the oxidation mechanisms of pharmaceuticals by PFA and UV-C/PFA and the pH influence, the TPs generated during the oxidation of LID and DCF in 10 mM PBS (pH 7.0) at five levels of pH ranging from 2.0–11.0 were characterized by high resolution mass spectrometry (HRMS). The exact masses of LID and DCF and their TPs, as well as their detection parameters and main fragments, are described in Tables S2 and S3, and the detailed identifications for each TP at various pH as well as the influence of quenchers (TBA and MeOH) on the TP formation are described in Text S5. A mechanism was proposed when the identification of the TP was possible (Figure S7).

As observed previously for PFA oxidation [46,48], lidocaine N-oxide ($m/z$ 251.1755) was the most intense TP under PFA and UV-C/PFA (Table S2). Its higher intensity in alkaline conditions can be related to the higher reactivity of LID in those conditions. For DCF oxidation, no TPs were observed at all pH values under PFA oxidation, likely due to the low oxidation rate of DCF by PFA [46], while seven TPs were

detected under UV-C/PFA, some of which corresponded to the loss of a chlorine atom or the carboxylic group (when the identification of the structure was possible, Table S2 and Figure S7). However, due to the inability to propose structures for all the observed TPs, it remains challenging to establish a direct correlation between their exclusive presence and the related pH conditions.

Overall, the TPs formed by PFA oxidation had higher molecular weights compared to those obtained under UV-C/PFA, probably due to the different oxidation mechanisms involved in the two processes and the enhanced oxidation rates obtained with AOPs. UV-C/PFA involves processes like homolytic bond dissociation, photocleavage, and ring cleavage induced by HO• [28,51,67] from the background UV-C/H₂O₂, leading to smaller molar mass TPs, while PFA primarily adds oxygen atoms, resulting in higher molar mass TPs [46].

### 3.4. Removal of pharmaceuticals by UV-C/PFA in wastewater effluent

To study the impact of UV-C/PFA on pharmaceutical degradation under realistic conditions, six pharmaceuticals were spiked into treated wastewater and then subjected to oxidation at 10 mg.L⁻¹ (161 µM) of PFA and 60 min of reaction. Prior to spiking, the analysis of the non-spiked wastewater samples (experiment conducted over three days) revealed the presence of these pharmaceuticals at varying concentrations ($C_0$), ranging from 0.3 µg.L⁻¹ for SMX to 1.3 µg.L⁻¹ for DCF. In one sample, FUR and ACT were detected at 3.4 µg.L⁻¹ and 16.7 µg.L⁻¹, respectively.

As expected (based on the $k_{UV-C/PFA}$ of pharmaceuticals compared to $k_{PFA}$, Table 1), UV-C/PFA increased the removal rates of all pharmaceuticals (Fig. 6), ranging from 10.9 ± 2.5 % for CBZ to 76.1 ± 5.5 % for DCF. These removal rates were consistent with the $k_{UV-C/PFA}$ results (part 3.3.2), which placed DCF among the most reactive compounds, CBZ as the least reactive compound and a slight difference (only 20.0 %) was
<figure><img src="image_8.png" /><figcaption>**Fig. 6.** The removal rate of pharmaceuticals in wastewater effluent by UV-C/PFA and PFA (data obtained from Nabintu Kajoka et al., [46]). Experimental conditions: [Pharmaceutical]₀ = C₀ + 1.0 µg.L⁻¹ spiked in wastewater effluent, with C₀ the initial concentration of the pharmaceutical in the wastewater effluent, reaction time = 60 min, pH = 7.9 ± 0.1, 20.0°C, [PFA]₀ = 10 mg.L⁻¹ = 161 µM, and UV-C fluence = 0.5 ± 0.01 mW.cm⁻² = 2.3 ± 0.03 × 10⁻⁷ Einstein.L⁻¹.s⁻¹. Error bars represent the standard deviation from triplicate samples.</figcaption></figure>

observed between the reactivity of LID under PFA and UV-C/PFA. For the oxidation by PFA, the maximum removal rate observed for these six compounds was 33.6 ± 13.6 % for FUR [46]. All the observed removal rates, however, were lower than expected. In kinetic experiments in PBS, four pharmaceuticals (DCF, SMX, ACT, and FUR) were completely degraded by 31 mg. L⁻¹ (500 µM) of PFA in less than 5 min. The fact that none of these pharmaceuticals were completely removed within 60 min when spiked in wastewater effluent indicates that, as previously observed for PFA [46], their removal rate was hindered by wastewater constituents, such as total suspended solids, dissolved organic matter, alkalinity or dissolved salts.

## 4. Conclusion

This study investigated, for the first time, the degradation kinetics of six pharmaceuticals by UV-C/PFA. A synergistic effect was observed between UV-C photolysis and PFA, significantly enhancing the removal rate of all pharmaceuticals in PBS or wastewater effluent compared to PFA alone. This improved degradation resulted from different reactive species, including ¹O₂, HCOO*, HCOOO* and HO*, with HO* originating principally from the background UV-C/H₂O₂. Although UV-C/PFA does not generate more HO* than UV-C/H₂O₂ (a common AOP), it offers an advantage due to the significant additional contribution of R-O* and ¹O₂ in the degradation of some pharmaceuticals (LID, FUR, SMX, and ACT). The contribution of reactive species to the degradation of pharmaceuticals was investigated, but supplementary quenching experiments with sodium azide and the determination of the steady-state concentration of ¹O₂ should be performed to distinguish the role of the PFA molecule from that of ¹O₂, as the latter appears to be the most reactive species in PFA oxidation. Nevertheless, the contribution of PFA to the degradation of PFA-reactive molecules (e.g., LID, FUR) under UV-C/PFA is lower than when PFA is used alone. Therefore, determining the optimal sequence for coupling the two processes is crucial to maximize their benefits. For instance, a first reactor could be used for PFA disinfection and reaction

with PFA-reactive molecules, followed by a second UV-C reactor in series to perform an AOP with the residual PFA for treating more recalcitrant pharmaceuticals. In wastewater effluent, UV-C/PFA increased the removal rate of all pharmaceuticals compared to PFA. However, the removal rates were lower than expected (based on their $k_{UV-C/PFA}$) due to scavenging effects from wastewater constituents such as total suspended solids, alkalinity, organic matter, or dissolved salts. The insights gained from this study provide a valuable foundation for further research on UV-C/PFA and other PFA-based AOPs. Future investigations could focus on the impact of wastewater constituents, as well as the toxicity of oxidized waters and TPs compared to other AOPs, which would advance the development of UV-C/PFA as a novel strategy for micropollutant removal in water treatment. Despite the promising results of UV-C/PFA, the addition of UV-C photolysis increases electrical consumption. Therefore, techno-economic assessments are necessary to demonstrate the feasibility and advantages of this novel AOP.

## Environmental implication

Performic acid (PFA) is an innovative disinfectant effective in removing pathogenic microorganisms, including *Escherichia coli*, enterococci, and viruses, from wastewater. However, due to its limited and selective reactivity with pharmaceuticals, PFA requires activation to target these contaminants effectively. In this study, PFA was combined with UV-C photolysis to create an advanced oxidation process (UV-C/PFA) that enhances pharmaceutical degradation by generating reactive species, such as hydroxyl and peroxyl radicals, and singlet oxygen. These results indicate the potential to extend PFA's role beyond disinfection, utilizing it as an oxidant to treat a broader range of emerging contaminants in wastewater.

## CRediT authorship contribution statement

**Diop Yacine Khadija:** Formal analysis. **Chebbo Ghassan:** Funding acquisition. **Gasperi Johnny:** Writing – review & editing, Conceptualization, Validation, Supervision. **Oliveira Marcos:** Validation, Supervision, Writing – review & editing, Conceptualization. **Vincent Rocher:** Funding acquisition. **Christelle Nabintu Kajoka:** Writing – original draft, Conceptualization, Investigation, Methodology. **Le Roux Julien:** Supervision, Conceptualization, Writing – review & editing, Validation. **Brosillon Stephan:** Validation, Conceptualization, Writing – review & editing, Supervision. **Reibel Corine:** Formal analysis.

## Declaration of Competing Interest

The authors declare that they have no known competing financial interests or personal relationships that could have appeared to influence the work reported in this paper.

## Acknowledgments

The authors thank the OPUR and WaterOmics (ANR-17-CE34-0009-01) research programs for their financial support. We are grateful to the PRAMMICS Platform (OSU-EFLUVE UMS 3563) for their assistance with UPLC-IMS-QTOF and HPLC-DAD analyses. We also thank Florian Huet for his assistance in building the UV-C photolysis system, as well as Emilie Caupos, Emmanuelle Mebold, Lila Boudhamane, Melissa Ben-Iken, and Ilyess Taibi for their contributions to various experiments and analyses. Finally, we thank Gabson Baguma (University of Nevada) for improving the clarity and quality of the manuscript.

## Appendix A. Supporting information

Supplementary data associated with this article can be found in the online version at doi:10.1016/j.jhazmat.2025.139016.
# Data availability

Data will be made available on request.

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