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Relevance of photocatalytic redox transformations of selected pharmaceuticals

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in a copper- and iron-rich Mediterranean intermittent river

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Maria Vittoria Barbieri* and Serge Chiron

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UMR HydroSciences Montpellier, University of Montpellier, IRD, CNRS, 15 Av. Charles Flahault 34093

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Montpellier cedex 5, France.

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*Corresponding author: maria-vittoria.barbieri@umontpellier.fr

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13 Abstract

This work aimed at investigating specific attenuation pathways of pharmaceuticals in copper- and iron-rich Mediterranean intermittent and sunlit rivers by combining lab- and field-scale studies. Poorly photodegradable and biodegradable compounds such as fluconazole, oxazepam and venlafaxine attenuated in two river stretches with short hydraulic residence times (< 3 h). This result was assumed to be related to their capacity to interact with photoreactive free Cu²⁺ and Fe³⁺ or their associated oxides. Lab-scale photodegradation experiments under simulated solar irradiation revealed the beneficial impact of a mixture Cu²⁺ and colloidal iron hydroxides at environmental concentrations and at neutral pH on the pharmaceuticals photodegradation kinetic rate constants. These latter were consistent with the in-stream attenuation rate constants of targeted contaminants which ranged from 0.104 to 0.154 h⁻¹. Further identification of phototransformation products by LC-HRMS highlighted reductive transformation pathways including reductive dehalogenation and hydrogenation reactions. Several TPs were found to be stable under irradiation and were detected in field monitoring, accordingly. This was ascribed to the formation of a Cu/Fe composite material under solar irradiation with photocatalytic properties. The role of Cu was to trap the electron in the conduction band of the iron-based photocatalyst, which promoted separation efficiency of electron-hole pairs as well as enhanced photoreduction processes at the expense of oxidation ones. Even though, these mechanisms have been reported in water treatment field for organic micropollutants removal, their significance was demonstrated for the first time in natural settings.

Keywords: Copper; in-stream attenuation; pharmaceuticals, photocatalytic reductive reactions; transformation pathways.
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# 1. Introduction

Attenuation of pharmaceutical mixture concentrations such as hydrochlorothiazide, valsartan and naproxen has mostly been studied in wastewater effluent-dominated perennial streams in temperate-regions and more scarcely in temporary rivers where increased pharmaceutical concentrations could be observed due to the lack of dilution capacity (Mandaric et al., 2019). These studies were mainly conducted by using a combination of field studies and controlled laboratory experiments to investigate the attenuation dynamics and the associated predominant mechanisms of pharmaceutical dissipation along river stretches. A main outcome of these studies has been that variable pharmaceuticals inputs and differential in-stream attenuation generate evolving complex mixtures along stream reaches during base flow conditions with unpredictable impacts to biota (Zhi et al., 2020). Major factors affecting the extent of in-stream attenuation intensity of pharmaceuticals and their transformation products (TPs) including sorption, photo- and bio-transformations are still a matter of discussion (Jaeger et al., 2019). However, the relevance of each process is compound specific and the in-stream attenuation rate is highly variable among river segments (Acuna et al., 2014).

The primary mechanism for cationic pharmaceuticals attenuation (e.g., venlafaxine and citalopram) is thought to be sorption to sediment and biofilm (Writer et al., 2013) while for more polar and/or anionic compounds which are expected to remain in the water phase (e.g., diclofenac and hydrochlorothiazide) photodegradation may prevail (Schmitt et al., 2021). Biodegradation may become relevant providing that the hydraulic retention time is long enough, since increased water travel time accounts for higher attenuation of most of pharmaceuticals (Mandaric et al., 2019). This was demonstrated for instance for metoprolol by implementing an enantiomeric fractionation methodology (Kunkle and Radke, 2012). However, most current understanding of chemical attenuation mechanisms in aquatic systems comes from laboratory experiments, which do not cover
all the environmental conditions and can yield different transformation rates and TPs than real-world conditions.

A special case is undoubtedly Mediterranean intermittent streams because the Mediterranean hydrological regime is characterized by brief and intense flow events after dry periods of small streams. This implies an important physical erosion associated with copper and iron transfer to rivers by suspended matter in water runoff (Xue et al., 2000) leading to iron- and copper-rich river sediment. For instance, high Cu concentration levels could be determined in river-bed sediments of a small Mediterranean vineyard catchment (60 - 176 µg/g) together with a high content of Fe₂O₃ (5.4%) as well as in river suspended matter (146 - 191 µg/g), while dissolved Cu fraction of river waters could reach up to 8 µg/L (El Azzi et al., 2013). High levels of Cu has frequently originated from high repeated spraying of Bordeaux mixture (i.e., Ca(OH)₂ + CuSO₄) for crops protection against fungal diseases. Those streams also receive high amounts of organic matter from domestic wastewater treatment plants (WWTPs) and undergo changes in redox conditions, including from wet-dry cycles resulting from variable flow regimes. Consequently, dissolved Fe and Cu is produced through reductive dissolution of Cu oxides and Fe oxy(hydr)oxides, driven by microbial degradation of organic matter (Iles et al., 2022).

We hypothesized that Cu and Fe alone or as metal composite might contribute to the in-stream attenuation of organic micropollutants because Cu and Fe complexes with organic ligands including organic contaminants are photoreactive (Sykora, 1997) and because iron and copper oxides as solid or colloidal particles have shown photocatalytic activities (Asif et al., 2021; Sibhatu et al., 2022). Their photocatalytic properties have been deeply implemented in water treatment for organic micropollutants removal. In contrast, their relevance in natural waters has been poorly investigated. This will be the major contribution of this work by combining a field study during two months in the Crieulon and Vidourle River watersheds and controlled laboratory experiments as recommended (Zhi et al., 2021). The Crieulon River is a small Mediterranean intermittent stream, tributary of the
Vidourle river, with only one-point source of contamination (i.e., a psychiatric hospital effluent discharge), making it a near-ideal field site to study the attenuation behaviors of pharmaceutical mixtures (see a map in Fig. S1).

Specific objectives included (i) a monitoring study of selected pharmaceuticals along two river stretches using passive samplers, (ii) photodegradation kinetic experiments under simulated sunlight irradiation using Cu²⁺/Fe³⁺ ions as well as different micrometric iron and copper oxides and spinel copper ferrite and (iii) TPs identification for selected compounds using liquid chromatography high resolution-mass spectrometry (LC-HRMS) for a deeper understanding of photodegradation mechanisms and their significance in investigated rivers. In this study, we will address and demonstrate the enhanced efficiency in photocatalytic oxidation/reduction of pharmaceuticals due to the formation of a Cu/Fe complex under solar irradiation for the first time in natural attenuation settings.

## 2. Materials and Methods

### 2.1 Chemicals

Analytical standards of targeted compounds atenolol, candesartan, carbamazepine, céliprolol, fluconazole, lamotrigine, oxazepam, sulpiride, tramadol, venlafaxine and their deuterated homologues, N,N-didesmethylvenlafaxine and venlafaxine-N-oxide were purchased from Toronto Research Chemicals (Toronto, Canada). Cetirizine, paroxetine and propanol were purchased from Sigma Aldrich (St Quentin-Fallavier, France). For extraction process and LC analysis, ultra-pure water (UPW) was obtained using a Millipore system; HPLC-grade acetonitrile (ACN) and methanol (MeOH) were purchased from Carlo Erba Reagents S.A.S. (Val de Reuil, France). Formic acid (FA) puriss p.a. ACS reagent analytical grade was purchased from Sigma-Aldrich (St Quentin-Fallavier, France). For the inorganic compounds, iron(II) oxide, iron(III) sulfate and spinel copper ferrite were purchased from Sigma-Aldrich (St Louis, USA). Copper(II) oxide was purchased from Alfa Aesar by Thermo Fischer Scientific (Ward Hill, USA), and copper(II) sulfate pentahydrate was purchased from Merck
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KGA (Darmstadt, Germany). Humic acids sodium salt and ascorbic acid were purchased from Sigma-
Aldrich (St Louis, USA).

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## 2.2 Study site

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Field surveys were conducted along two river stretches located in the northern part of Montpellier
city (France). The first one of 1.7 km was selected on the Crieulon River with two sampling sites (SP1
and SP2, see Fig. S1) and the second one of 5.4 km on the Vidourle River with two sampling sites (SP3
and SP4). SP3 was located just after the confluence of the Crieulon River with the Vidourle River so
that some dilution was expected. The Crieulon and the Vidourle are two small Mediterranean
catchments, which were mainly occupied by vineyard, with a prolonged dry season and a mean
annual average rainfall of 620 mm. The first stretch received water from one single continuous
wastewater point-source from an effluent discharge of a psychiatric hospital (WWTP, SP1) while the
second one received treated wastewater from different small villages located upstream to SP3. The
WWTP outfalls were considered the principal sources of pharmaceuticals with minimal expected
non-point source, representing a suitable study site for investigating the fate of these chemicals in
surface water.

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## 2.3 Sampling procedure

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A tracer test was conducted the day before the sampling to estimate hydrologic travel times using
fluorescent dye (rhodamine WT). Travel times from SP1 to SP2 and from SP3 to SP4 averaged 1.8 h
and 2.6 h, respectively. River flow rates at the different SP were measured using an acoustic Doppler
velocity meter (ADV; FlowTracker, SonTek, San Diego, CA, U.S.A.). The recorded values were 2.3, 2.3,
4.2 and 4.3 m³s⁻¹ at SP1, SP2, SP3 and SP4, respectively. Information on the quality of river waters at
the 4 sampling points is reported in Table S1. High concentrations in dissolved Fe and Cu, in the 250 -
300 and 2 - 5 µg/L range, respectively, were the main feature of these waters (see Table 1). Polar
organic chemical integrative samplers (POCIS) were used for environmental sampling because they
have been demonstrated to be suitable to determine the attenuation of organic micropollutants in
river and lead to similar results as those obtained through high-resolution sampling (e.g., Lagrangian
sampling scheme; Jaeger at al., 2019; Li et al., 2016). The targeted compounds concentrations in ng/L in river water were not needed for attenuation assessment because it was assumed that the uptake kinetics of a compound were identical along each river stretch, which was a reasonable assumption as the river characteristics did not change substantially between SP1 and SP2 as well as between SP3 and SP4. Consequently, the compounds amount was expressed in ng/g of dry POCIS sorbent. There was no field blank for POCIS, but the results from the procedural blank samples showed that there was no contamination during sample extraction and analysis.

Environmental sampling was carried out in June and July 2021 and POCIS were let for pharmaceuticals accumulation during three weeks. At each sampling site, one cage containing 3 POCIS discs was vertically submerged with a total sampling area of 41 cm². AttractSPE®POCIS HLB were purchased from Affinisep (Houlme en Normandie, France) consisting of approximately 230 mg of the solid adsorbent N-vinylpyrrolidone-divinylbenzene (Oasis HLB), which is a universal sorbent used to extract a wide range of compounds with a large range of polarity (Alvarez, 2010). During the sampling period, weather was mostly sunny with only two clouded days and no rain. In these conditions, solar radiation intensity ranged between 200 Wm⁻² during clouded days to 1000 Wm⁻² at midday during the sunny days.

Compounds attenuation rate constants $k_{att}$ along the two river sections were determined assuming a first-order process from the slope of Eq. 1 (Ln[C]t/[C]t₀ = $k_{att}$ x t), with [C]₀ at the sampling sites 1 and 3 and [C]ₜ at the respective downstream sites 2 and 4. The validity of this approach assumed no major dilution or input of the chemicals along the river sections and was tested under the conditions during our field study with carbamazepine, which is known as a conservative contaminant (Jaeger at al., 2019). Carbamazepine showed negligible attenuation along the two river stretches (see Table 1). Compounds $k_{att}$ were determined based on a mean travel time of 1.8 h from SP1 to SP2 and 2.6 h from SP3 to SP4.

## 2.4 Photochemical studies
Photolysis of target compounds was carried out in parallel experiments in Pyrex glass cells with distilled water spiked with 10 mg/L of each compound (direct photolysis) and with 10 mg/L of micrometric iron oxide, copper oxide, spinel copper ferrite, Cu²⁺, Fe³⁺, and Cu²⁺/Fe³⁺ mixture (in separate experiments). In the Cu²⁺/Fe³⁺ experiment, [Cu²⁺] and [Fe³⁺] was 10 µg/L each as well as the concentrations of pharmaceuticals (10 µg/L) in presence of 10 mg/L of humic substances to better reflect environmental conditions. The irradiation was performed using a 1500 W Xenon lamp (Suntest, CO.FO.MEGRA, Italy) equipped with a 290 nm cut-off filter simulating solar light. The irradiance was set at 750 W m⁻² because in Montpellier area, in summer time (sampling period), solar irradiance can reach 1000 Wm⁻² at midday during a sunny day. The water temperature reached during the irradiation was 55°C and the pH was adjusted in each experiment with the addition of sodium hydroxide (NaOH) if necessary to remain constant at 7.5. This pH value was very close to the ones of river water samples. A temperature of 55°C was not realistic but it is known that photochemical kinetics do not significantly depend on temperature.

One mL of aliquots was collected at pre-determined intervals over the course of 24 h. The remaining water at the end of each test was collected and the volume measured to determine loss via evaporation for adjustment. Each experiment was conducted in triplicate and results were reported as mean values. Iron and copper oxide dissolution with ascorbic acid was performed based on a methodology proposed by Joseph et al. (1995) to confirm sorption processes of pharmaceuticals. All samples were filtered through 0.45 µm syringe filters and directly analyzed by LC-HRMS for kinetic studies and TPs identification. The phototransformation rate constants $k_{\text{photo}}$ and related half-lives ($T_{1/2}$) were determined analogously to Eq. 1 ($\ln[C]_t/[C]_{t0} = k_{\text{photo}} \times t$) with $[C]_t$ referring to concentrations of compounds at each time point relative to the starting concentration $[C]_{t0}$.

## 2.5 Sample processing and analysis

**Target analysis:** Upon retrieval, POCIS discs were individually washed with distilled water, sealed in its original aluminum bag, transported to the lab in ice box and stored under -20°C until extraction.
Receiving phases of the POCIS were transferred into 3 mL empty polypropylene SPE cartridges, spiked with deuterated standards (1 mg/L acetonitrile), and then eluted with 8 mL of acetonitrile. After concentration under a gentle stream of nitrogen, final extracts (1.5 mL) were analyzed by LC-HRMS. The chromatographic run was carried out by means of Dionex™ Ultimate 3000 UHPLC (Thermo Scientific™, San Jose, USA) using a Waters XBridge C₁₈ column (150 x 2.1 mm i.d., 2.5 µm particle size). The mobile phases were 0.1% FA in water (A) and 0.1% FA in MeOH (B) under the following gradient: 95%A:5%B from 0 to 1 min, 90% B at 19 min with a 1 min hold, 5% B at 20.5 min, followed by a 5 min equilibration at 95%A:5%B. The flow rate was 0.3 mL/min. The MS analysis was performed with an Q-Exactive™ Focus Orbitrap mass spectrometer (Thermo Fisher Scientific, Les Ullis, France) equipped with an HESI electrospray source and operated in positive ionization mode in full scan acquisition. Further details of MS and HESI parameters are provided in SM. For the quantification of targeted pharmaceuticals, matrix calibration curves were made according to the matrix match standard procedure using the internal standard method (European Commission, 2017). Analytical limits of detection (LODs) and limits of quantifications (LOQs), as well as linearity, accuracy and precision of the method have been previously validated in our group and reported in Tadić et al. (2022).

*Non-target analysis:* Samples were analyzed using a Vanquish ultra-high performance liquid chromatography (Thermo, Fisher Scientific, Waltham, MA, USA) coupled to a high-resolution mass spectrometry Q-Focus Orbitrap (Thermo Fisher Scientific, Waltham, MA, USA) equipped with a heated electrospray ionization (HESI) source operated in full scan + data dependent acquisition MS2 (DDA-MS2) mode. The chromatography assay was performed using 10 µL injection volume, 0.3 mL/min flow rate, and a binary gradient of water (A) and acetonitrile (B), both containing 0.1% formic acid, as follows: 10% B at 0–1 min, 90% B at 10–20 min, 10% B at 21–27 min. Data were treated using Thermo Xcalibur™ 3.1 Software (Thermo Fisher Scientific, CA, USA). The specific workflow and instrumental conditions used for elucidating transformation products is provided in SM.
<table><caption>**Table 1.** Concentrations in ng/g of POCIS sorbent ± relative standard deviation (RSD) for targeted pharmaceuticals and in µg/L for dissolved Cu and Fe at four sampling sites (SP1, SP2, SP3 and SP4) along the investigated river stretch during June and July 2021. See the map in Supporting Material for sampling sites location.</caption><thead><tr><th>Compound</th><th>SP1</th><th>SP2</th><th>SP3</th><th>SP4</th></tr></thead><tbody><tr><td colspan="5" style="text-align: center; font-weight: bold;">June 2021</td></tr><tr><td>Sulpiride</td><td>&lt;LOQ</td><td>&lt;LOQ</td><td>1.9 ± 0.4</td><td>1.6 ± 0.35</td></tr><tr><td>Atenolol</td><td>&lt;LOQ</td><td>&lt;LOQ</td><td>2.3 ± 0.3</td><td>0.8 ± 0.1</td></tr><tr><td>Tramadol</td><td>43.8 ± 5.3</td><td>34.3 ± 4.1</td><td>27.0 ± 3.2</td><td>23.4 ± 2.8</td></tr><tr><td>Fluconazole</td><td>379.2 ± 45.5</td><td>265.7 ± 29.5</td><td>172.3 ± 20.7</td><td>103.8 ± 12.5</td></tr><tr><td>Lamotrigine</td><td>348.5 ± 38.3</td><td>326.5 ± 35.9</td><td>56.8 ± 6.2</td><td>61.4 ± 6.8</td></tr><tr><td>Celiprolol</td><td>20.7 ± 1.7</td><td>14.7 ± 1.2</td><td>37.3 ± 3.0</td><td>15.2 ± 1.2</td></tr><tr><td>Venlafaxine</td><td>175.5 ± 14.0</td><td>129.3 ± 10.3</td><td>59.2 ± 4.7</td><td>36.8 ± 2.9</td></tr><tr><td>Propanolol</td><td>7.8 ± 0.5</td><td>6.7 ± 0.4</td><td>143.4 ± 8.6</td><td>66.2 ± 4.0</td></tr><tr><td>Paroxetine</td><td>6004.7 ± 420.3</td><td>3766.5 ± 263.7</td><td>1932.2 ± 135.3</td><td>730.9 ± 51.2</td></tr><tr><td>Cetirizine</td><td>8.7 ± 1.2</td><td>4.7 ± 0.7</td><td>7.2 ± 1.0</td><td>5.8 ± 0.8</td></tr><tr><td>Oxazepam</td><td>12244.5 ± 979.6</td><td>8948.3 ± 795.9</td><td>3199.3 ± 255.9</td><td>1632.3 ± 130.6</td></tr><tr><td>Candesartan</td><td>2.1 ± 0.2</td><td>2.9 ± 0.2</td><td>5.3 ± 0.4</td><td>4.6 ± 0.4</td></tr><tr><td>Carbamazepine</td><td>224.6 ± 11.2</td><td>232.7 ± 11.6</td><td>188.4 ± 9.4</td><td>182.6 ± 9.1</td></tr><tr><td>Soluble Cu</td><td>4.834 ± 0.145</td><td>3.679 ± 0.110</td><td>5.745 ± 0.172</td><td>2.265 ± 0.098</td></tr><tr><td>Soluble Fe</td><td>255.4 ± 10.2</td><td>324.2 ± 13.0</td><td>288.3 ± 11.5</td><td>312.5 ± 12.5</td></tr><tr><td colspan="5" style="text-align: center; font-weight: bold;">July 2021</td></tr><tr><td>Sulpiride</td><td>&lt;LOQ</td><td>&lt;LOQ</td><td>4.0 ± 0.9</td><td>3.6 ± 0.8</td></tr><tr><td>Atenolol</td><td>&lt;LOQ</td><td>&lt;LOQ</td><td>1.7 ± 0.2</td><td>&lt;LOQ</td></tr><tr><td>Tramadol</td><td>12.6 ± 1.5</td><td>10.3 ± 1.2</td><td>178.6 ± 21.4</td><td>156.7 ± 18.8</td></tr><tr><td>Fluconazole</td><td>280.4 ± 22.4</td><td>221.7 ± 17.7</td><td>132.6 ± 10.6</td><td>85.4 ± 6.8</td></tr><tr><td>Lamotrigine</td><td>245.6 ± 27.0</td><td>224.2 ± 24.7</td><td>76.8 ± 8.4</td><td>65.6 ± 7.2</td></tr><tr><td>Celiprolol</td><td>24.5 ± 2.0</td><td>18.4 ± 1.5</td><td>61.9 ± 5.0</td><td>19.4 ± 1.6</td></tr><tr><td>Venlafaxine</td><td>145.6 ± 11.6</td><td>110.6 ± 8.8</td><td>48.7 ± 3.9</td><td>32.4 ± 2.6</td></tr><tr><td>Propanolol</td><td>&lt;LOQ</td><td>&lt;LOQ</td><td>4.9 ± 0.3</td><td>&lt;LOQ</td></tr><tr><td>Paroxetine</td><td>4564.8 ± 319.6</td><td>3067.3 ± 214.7</td><td>2568.2 ± 179.8</td><td>1120.5 ± 78.4</td></tr><tr><td>Cetirizine</td><td>6.4 ± 0.9</td><td>5.3 ± 0.7</td><td>9.2 ± 1.3</td><td>8.5 ± 1.2</td></tr><tr><td>Oxazepam</td><td>8934.6 ± 714.8</td><td>5896.4 ± 471.7</td><td>2435.8 ± 194.9</td><td>1296.6 ± 103.7</td></tr><tr><td>Candesartan</td><td>2.6 ± 0.2</td><td>2.2 ± 0.2</td><td>37.5 ± 3.0</td><td>35.8 ± 2.9</td></tr><tr><td>Carbamazepine</td><td>318.7 ± 15.9</td><td>312.3 ± 15.6</td><td>247.4 ± 12.4</td><td>236.9 ± 11.8</td></tr><tr><td>Soluble Cu</td><td>4.245 ± 0.127</td><td>3.456 ± 0.104</td><td>5.423 ± 0.163</td><td>2.867 ± 0.086</td></tr><tr><td>Soluble Fe</td><td>232.4 ± 9.9</td><td>312.2 ± 12.5</td><td>284.3 ± 11.4</td><td>308.5 ± 12.3</td></tr></tbody></table>

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214 3. Results and discussion

215 3.1 Occurrence and attenuation of targeted pharmaceuticals along two river stretches

216 For pharmaceuticals monitoring, a first LC-HRMS-based non-target screening workflow was implemented for the identification of 47 pharmaceuticals in river water samples (Tadić et al., 2022) of which 13 compounds were commonly quantified in POCIS extracts mainly including β-blockers (3), and psychiatric drugs (6) (see Table 1). The most persistent compounds were tramadol, lamotrigine,
220 sulpiride, candesartan and carbamazepine while the other ones including atenolol, céliprolol,
221 fluconazole, venlafaxine, propranolol, paroxetine, and oxazepam underwent dissipation to different
222 extents along the river reaches. To check for hydraulic dilution, a conservative tracer (i.e.,
223 carbamazepine) was used. Carbamazepine concentrations remained stable between sampling sites
224 SP1 and SP2 as well as between SP3 and SP4 (see Table 1). Consequently, a decrease in
225 concentrations between two consecutive sampling sites could only be ascribed to sorption, photo-
226 and/or biotransformation processes and not to dilution. The riverbeds consisted mainly of gravel and
227 sand with a lack of macrophytes so that absorption of contaminants by plants could be ruled out.
228 However, their absorption by microorganisms and algae was likely to occur.

229 To gain evidence if biotransformation has occurred, the potential enantiomeric fractionation (EF) for
230 chiral pharmaceuticals has been applied (Mechelke et al., 2020). Venlafaxine was selected for this
231 purpose because biotransformation has been reported as an important in-stream attenuation
232 process for venlafaxine (Schmitt et al., 2021). EF between SP1 and SP2 and between SP3 and SP4
233 were determined for selected samples with chiral chromatography following a methodology
234 previously established by our group (Li et al., 2013). No significant change in EF was observed (see
235 Fig. S2). EF is known to be strongly dependent on environmental conditions (e.g., microbial
236 communities, Mechelke at al., 2020) most likely accounting for discrepancies between the results of
237 this study and those reported by Schmitt et al. (2021). The absence of EF of venlafaxine could not
238 fully exclude biotic processes for other compounds because potentially different enzymes are
239 responsible for the transformation of different drug families. However, biotransformation could be
240 reasonably discarded as significant dissipation pathways at least for poorly biodegradable
241 compounds such as fluconazole and oxazepam (Pacholak et al., 2022; Redshaw et al., 2008).

242 Sorption to sediment and photodegradation were very likely the two major processes accounting for
243 pharmaceuticals attenuation. The behavior of some compounds was anticipated such as that of
244 paroxetine due to quick direct and indirect photolysis (Gornick et al., 2021) and that of venlafaxine
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and β blockers through sorption to sediment because they are positively charged at environmental
pH facilitating interaction with negative surfaces of sediment (Li et al., 2015). To confirm this
assumption, analysis of sediment samples was performed for selected compounds (see analytical
methodology and results in Table S2 and S3).

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Lamotrigine, tramadol, celirolol, venlafaxine, propanolol and citalopram, which are basic drugs and
positively charged at neutral pH were all quantified at a few ng/g levels. Oxazepam, a neutral
compound but with a log K<sub>ow</sub> = 2.2 was also detected in the 5 - 7 ng/g range while fluconazole,
another neutral compound but with a lower K<sub>ow</sub> = 0.5, was never detected. The fast attenuation of
fluconazole was unexpected because this latter has frequently been used as a conservative tracer in
in-stream attenuation studies due to its known high stability under photolysis and under
biodegradation (Li et al., 2015) and lack of affinity for sediment. Fluconazole quickly dissipated with a
mean attenuation rate constants (k<sub>att</sub>) of 0.1 h<sup>-1</sup> over the two sampling campaigns in June and July,
assuming a pseudo first-order reaction (see experimental section). However, dissolved Fe and Cu
heavily occurred in the investigated waters with concentrations in the 2.3 - 5.7 and 255.4 – 324.2
µg/L range, respectively as well as in river sediments (see Table 1). Fluconazole is known for its
capacity to form complexes with transition metals such as Fe and Cu (Stevanović et al., 2021).
Consequently, these two transition metals were assumed to play a role in fluconazole in-stream
attenuation rates due to the potential photoreactivity of such complexes. This might hold true for
oxazepam (Correira dos Santos et al., 2002) and venlafaxine (Alturiqui, 2018). In spite of some
sorption to sediment, their fast dissipation rates were not anticipated since oxazepam and
venlafaxine are not fast photodegradable compounds, with half-lives reported to be 66 and 58 h
(West and Rowland, 2012; Manasfi et al., 2022). In contrast, lamotrigine is not known to undergo
complexation with transition metals which might account for its persistence. Tramadol has been
reported to form complexes with Cu but remained stable. However, the stability constant of
tramadol complex with Cu (log β = 4.0, Mansoor and Farooqui, 2017) is less than half that of
venlafaxine (log β = 9.6, Alturiqi et al., 2018). To confirm the assumption of the involvement of Cu

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and/or Fe in the dissipation of selected compounds, lab-scale photodegradation experiments were conducted.

## 3.2 Photodegradation of selected pharmaceuticals under controlled lab conditions

Lab-scale experiments were conducted to study phototransformation separately from other processes (i.e., biotransformation and sorption) for a better understanding of in-stream attenuation mechanisms of selected compounds. For this purpose, three pharmaceuticals, i.e., oxazepam, venlafaxine and fluconazole were selected according to (i) their high occurrence in the investigated river samples, as well as (ii) their known slow direct photodegradation kinetics and (iii) their tendency to form complexes with transition metals (e.g., Cu(II) and/or Fe(II)). Fig. 1 shows photodegradation kinetics under different conditions using dissolved Cu²⁺ and Fe³⁺ for their potential ability to form complexes with targeted compounds, and Fe₂O₃ (hematite), CuO and spinel copper ferrite for their photocatalytic behavior (Fig. 1a-c). Experiments with equal amounts of Cu²⁺ and Fe³⁺ (i.e., 10 µg/L) were also carried out in presence of humic substances to better reflect environmental conditions (Fig. 1d).
<figure><img src="image_1.png" /><figcaption>**Fig. 1.** Photodegradation of oxazepam (a), fluconazole (b) and venlafaxine (c) spiked at 10 mg/L, and 10 mg/L of iron oxide, copper oxide, Cu²⁺, Fe³⁺, spinel copper ferrite, and (d) a detail of photodegradation kinetics of the three compounds at environmental concentrations (10 µg/L) in presence of Cu²⁺ and Fe³⁺ ions mixture at 10 µg/L, with the addition of 10 mg/L humic acid.</figcaption></figure>

No significant losses of oxazepam, venlafaxine or fluconazole were observed in the dark controls, eliminating the possibility of thermal or hydrolytic degradation. Sorption processes of pharmaceuticals mainly occurred with iron and copper oxides and colloidal iron hydroxides particles resulting from the poor solubility of Fe³⁺ at neutral pH occurred at T₀. This was confirmed by iron and copper oxide dissolution (see experimental section), which prompted the release of the sorbed fraction excluding others abiotic transformation pathways than the photochemical one. Results obtained under direct photolysis were consistent with those reported in literature, that is slow degradation kinetic rates for oxazepam and fluconazole and faster ones for venlafaxine (Rua-Gomez and Püttmann, 2013; Castro et al., 2016; Calisto et al., 2011).
The photodegradation of each compound followed rather well apparent first order kinetics ($r^2$ > 0.98) under different conditions and the related kinetic rate constants ($k_{photo}$) and half-lives ($T_{1/2}$) were calculated accordingly (see Table S4). The photodegradation kinetics were always faster with iron than with copper. The generation of hydroxyl radicals under the photolysis of the colloidal ferric hydroxide particles or under hematite photocatalytic activity most likely accounted for this result (Sun et al., 2014). The potential complexation of pharmaceutical with $Cu^{2+}$ did not result in significant photodegradation kinetics enhancement (in comparison to direct phototransformation) and even in deceleration in case of fluconazole and venlafaxine most likely due to light screening. This is in contrast with cephalosporines β-lactam antibiotics that underwent fast photodegradation in presence of $Cu^{2+}$ but in this case, a photoinduced hydrolysis of the lactam ring could explain the fast photoreactivity (Zhang et al., 2020).

When $Cu^{2+}$ was added to $Fe^{3+}$ aqueous solutions at environmental concentrations, an increase in $k_{photo}$ was observed for the three target compounds (see Fig. 1d and Table S4). In this specific case, the values of $T_{1/2}$ in lab and field studies agreed reasonably well considering the complexity of the investigated field system and the assumptions that were made (e.g., constant mean travel time for the whole sampling period). This holds particularly true for oxazepam with average $T_{1/2}$ of 4.5 and 5.5 h in the field and lab study, respectively (see Fig. 2).
<figure><img src="image_2.png" /><figcaption>**Fig. 2.** Comparison of half-lives (T₁/₂) of targeted compounds in field and lab-studies under photolysis in distilled water with 10 µg/L Cu²⁺ and Fe³⁺.</figcaption></figure>

Higher T₁/₂ values in the field than in the lab for venlafaxine might be attributed to partial adsorption to the sediment, the adsorbed fraction becoming less accessible to solar irradiation. Indirect photolysis of venlafaxine in river was likely to occur which has mainly been attributed to the formation of hydroxyl radicals (Rua-Gomez and Püttmann, 2013), contrary to oxazepam and fluconazole for which indirect photolysis was found insignificant (Calisto et al., 2011; Castro et al., 2016). This meant that in presence of Cu²⁺ and Fe³⁺, this pathway is likely becoming negligible. Lower T₁/₂ in the field than in the lab might be related to the insignificance of direct photolysis of fluconazole. The photocatalytic property of iron oxy(hydr)oxides is limited due to the fast recombination of electron-hole pairs. However, this photocatalytic activity in visible light is usually enhanced in the presence of transitional metals such as Cu²⁺. The role of this latter species is to trap the electron in the conduction band of the photocatalyst, which promotes separation efficiency of
334 electron-hole pairs as well as enhances photoreduction processes (e.g., dehalogenation) at the
335 expense of oxidation ones (Baumanis et al., 2011; Lei et al., 2016). However, this shift in
336 contaminant transformation pathways, if rather known in water treatment settings, has not been
337 demonstrated in natural attenuation settings. Photocatalysis is an interfacial reaction in which
338 pollutants will be degraded more efficiently when they diffuse to the surface of the catalysts. Several
339 studies have therefore shown that the synergy of adsorption and photocatalysis can efficiently
340 enhance the pollutants removal rates (Yu et al., 2021). The affinity of Cu and Fe to bind with
341 investigated pharmaceuticals might account for the observed enhanced efficiency in their
342 photocatalytic oxidation/reduction.

343 3.3 Identification of TPs in lab-scale experiments and their occurrence in river water

344 For a possible elucidation of the mechanisms of phototransformation of targeted compounds, we
345 proceeded to TPs identification under different experimental conditions using suspect and non-
346 target LC-HRMS screening.

347 For venlafaxine, a total of nine TPs were identified. N,N-didesmethylvenlafaxine and venlafaxine-N-
348 oxide were confirmed by comparison of the mass spectra and retention times to those of purchased
349 synthetic compounds. Two TPs were identified by comparison with literature spectra (TPven-294 and
350 TPven-292, Osawa et al., 2019). Five TPs were detected for the first time, the structures of which
351 were assigned by interpreting MS data (see Fig. S3 for extracted ion chromatograms (EIC) and MS²
352 spectra), mass accuracy, retention time plausibility, fragments ions, and on the basis of reactivity
353 knowledge.

354 As a matter of example, in Fig. 3 the fragmentation pathway of TPven-251 is proposed. First,
355 TPven251 has an even Mw indicating the loss of the nitrogen atom from venlafaxine structure. The
356 MS² spectrum shows two major fragment ion at m/z 152.0467 and 163.0752. These ions were
357 characteristic of cyclohexanol ring losses and the experimental masses at 152.0467 and 163.0752
358 could only match to C₉O₂H₁₂ and C₁₀H₁₁O₂ elemental composition, respectively, supporting the
359 formation of a primary alcohol functional group.

360 TPven-249, TPven-251 and TPven-219 most likely arose from a one-electron oxidation at the C in
361 position α of the secondary amine moiety leading to an unstable radical species. Even Mw of these
362 TPs suggested N losses. Consequently, this unstable radical could evolve in different paths, including
363 i) oxidation into an aldehyde following the formation of a hydroperoxyl radical in presence of O₂
364 (TPven-249), ii) reduction into the corresponding primary alcohol (TPven-251) and iii) neutralization
365 by dehydration and rearrangement (TPven-219). DBE values of TPven-251 (4.5) and TPven-253 (3.5)
366 differed in one unit, suggesting that TPven-253 most likely resulted from hydrogenation reaction (1
367 x2H) at the aromatic ring. An additional TP was detected at m/z [M+H]⁺ = 282.2790. DBE value of this
368 TP (2.5) was two units lower than that of venlafaxine (4.5) and hydrogenation reactions (2 x 2H) at
369 the venlafaxine benzene ring was assumed to lead to TPven-282 (see Table S5 for MS data and
370 proposed data and Fig. S3 for EIC and MS² spectra). No structure could be definitively assigned to
371 TPven-279.
<figure><img src="image_3.png" /><figcaption>**Fig. 3.** Total Ion Chromatogram (TIC) with mass spectral peaks belonging to TPs from photolysis experiments with a detail of TPven-251 peak at RT= 13.15, and ESI full scan mass spectra (m/z 50-250) of degradation product TPven-251 with relative molecular ions and proposed fragmentation pathway.</figcaption></figure>

In the case of oxazepam, a total of five TPs were identified by comparison of literature spectra (Yang et al., 2018; West and Rowland, 2012) including TPoxa-257, TPoxa-241, TPoxa-273, TPoxa-223 and TPoxa-301 and two TPs remained unknown including TPoxa-227 and TPoxa-327 (see Table S5 for experimental [M+H]$^+$ accurate M<sub>w</sub>, possible elemental composition, together with mass error and double bond equivalents (DBE) and proposed structures). The photodegradation pathway under direct photolysis (blue arrow) was consistent with that previously reported with the formation of quinazoline (TPoxa-241) and quinazolinone (TPoxa-257) derivatives. The phototransformation pathway under Fe$^{3+}$ irradiation (green arrow) involved oxidation reactions through hydroxylation
386

(TPoxa-273 and TPoxa-301) while Cu²⁺/Fe³⁺ irradiation (orange arrow) accounted for a specific reduction reaction through reductive dechlorination leading to TPoxa-223.

387

388

Regarding fluconazole, five TPs were detected. The structures of TPflu-287 and TPflu-305 were confirmed thanks to the availability of literature spectra (Castro et al., 2016). Fluconazole dehydration (- H₂O) likely accounted for TPflu-289 formation. TPflu-301 and TPflu-317 were detected for the first time to the better of our knowledge. Their EIC together with their MS² spectra are shown in Fig. S3. As an aryl halide, TPflu-305 could undergo a radical-nucleophilic aromatic substitution through an aryl radical intermediary species following fluorine ion losses (Castro et al., 2016). This latter radical located in (m) position with respect to OH substituent shifted to (p) position by resonance accounting for the quinone formation (TPflu-301). TPflu-301 could evolve into TPflu-317 by further hydroxylation. The chemical structures of TPflu-269 and TPflu-257 could not be elucidated. On the basis of TPs identification, the suggested photodegradation fates of oxazepam, venlafaxine and fluconazole under different experimental conditions (i.e, direct photolysis, in presence of Fe³⁺ alone or a mixture of Fe³⁺/Cu²⁺) are presented in Fig. 4a, 4b and 4c, respectively.

389

390

391

392

393

394

395

396

397

398

399

a)

<figure><img src="image_4.png" /><figcaption>TPoxa-301</figcaption></figure>

<figure><img src="image_5.png" /></figure>

<figure><img src="image_6.png" /><figcaption>Oxazepam</figcaption></figure>

<figure><img src="image_7.png" /><figcaption>TPoxa-257*</figcaption></figure>

<figure><img src="image_8.png" /><figcaption>TPoxa-241*</figcaption></figure>

<figure><img src="image_9.png" /></figure>

<figure><img src="image_10.png" /><figcaption>TPoxa-223</figcaption></figure>

<figure><img src="image_11.png" /></figure>

<figure><img src="image_12.png" /><figcaption>TPoxa-273</figcaption></figure>

400
<figure><img src="image_13.png" /><figcaption>b)</figcaption></figure>

401

<figure><img src="image_14.png" /><figcaption>c)</figcaption></figure>

402

<caption>**Fig. 4.** Proposed photodegradation pathways of oxazepam, venlafaxine and fluconazole under (a) direct photolysis (blue arrows), (b) in presence of Fe³⁺ (green arrows) and (c) Fe³⁺ with Cu²⁺ (orange arrows).</caption>

406

* compounds found in real samples.

407

As far as the venlafaxine photodegradation pathway is concerned, the competition between oxidative and reductive pathways was not so noticeable between Fe³⁺ and Cu²⁺/Fe³⁺ systems.

408

409
However, Fe³⁺ system appeared more oxidative than Cu²⁺/Fe³⁺ system since N-dealkylation reactions leading to the formation of N, N-didesmethylvenlafaxine could be specifically attributed to Fe irradiation while Cu²⁺/Fe³⁺ irradiation was responsible for hydrogenation reaction leading to the conversion of TPven-251 into TPven-253 and the conversion of venlafaxine into TPven-282. A common path between direct photolysis and the Fe³⁺ and Cu²⁺/Fe³⁺ systems, initiated by a one-electron oxidation reaction at the C located in α position of the N atom was also observed. This reaction was enabled because the resulting radical was stabilized by resonance and could be further reduced, oxidized or neutralized to give TPven-219, TPven-249 and TPven-251. Other already known direct photodegradation pathways involved the formation of N-oxidized, hydroxylated (TPven-294) and ketone (TPven-292) derivatives.

Regarding fluconazole, the formation of TPflu-305 and TPflu-301 went through radical-nucleophilic substitutions of fluorine atoms. This is a reduction reaction with the formation of a radical anion as intermediary species. Fluorine is known to be a better leaving group than chlorine and defluorination could therefore also occur under direct photolysis. Why this substitution reaction did not proceed in Fe³⁺ system was not elucidated. In contrast, Fe³⁺ system led to a cyclisation reaction rather than the substitution reaction after the formation of the radical anion as well as to venlafaxine dehydration. The evolution of TPs abundances against time was followed on the basis of peak area to anticipate which TPs may accumulate in Cu²⁺/Fe³⁺ system (see Fig. 5).
<figure><img src="image_15.png" /><figcaption>**Fig. 5.** Evolution of TPs concentrations expressed in relative peak abundance against time after Cu²⁺/Fe³⁺ irradiation at environmental conditions (i.e. Cu²⁺/Fe³⁺ at 10 µg/L with the addition of 10 mg/L humic acid).</figcaption></figure>

All significant TPs were quickly formed during the first 60 min of irradiation and later on reached steady-state concentrations (i.e., formation and degradation offset one another) with the exception of hydroxyl-venlafaxine TP and TPoxa-223 which slowly underwent further degradation. This 1 h time frame of formation was compatible with TPs in-stream formation since HRTs were higher than 1 h and was consistent with the detection of a few of these TPs in field monitoring including TPoxa-257, TPoxa-241, venlafaxine-N-oxide, N,N-didesmethylvenlafaxine, hydroxy-venlafaxine, TPven-282, TPven-250, TPven-219, TPven-253, TPflu-289 and TPflu-301. Interestingly, while venlafaxine-N-oxide, hydroxyl-venlafaxine and N,N-didesmethylvenlafaxine already occurred in the hospital WWTP discharges (Manasfi et al., 2022), the other detected TPs were in-stream generated, highlighting the relevance of photocatalytic redox processes in sunlit river, especially reductive dehalogenation or hydrogenation reactions with the detection of TPven-282, TPven-253 and TPflu-301.
444 Co-catalyst-assisted photocatalytic reduction have been widely demonstrated using instance Rh-
445 loaded TiO₂ for hydrogenation of aromatic ring (Nakanishi et al., 2018) or CuO-modified TiO₂
446 nanocomposites for dehalogenation of polybrominated diphenyl ethers (Lei et al., 2016). In this latter
447 system, the supposed mechanism involved CuO trapped electrons from visible light irradiated TiO₂ to
448 form Cu₂O. The higher reduction potential of Cu₂O over CuO and TiO₂ allows for transferring
449 electrons to contaminants initiating the photocatalytic process. Similar behavior in the Cu²⁺/Fe³⁺
450 system might be anticipated where in irradiated solutions Cu²⁺ might be bound to the surface of the
451 hydrous iron oxide and/or CuO might precipitate at the surface of the oxide (Boukhalfaa et al., 2010).
452 This system to be effective required compound adsorption / complexation at the catalyst surface.
453 This explained what it only worked for those pharmaceuticals (i.e., fluconazole, oxazepam and
454 venlafaxine) which can form complexes with transition metals such as Cu and Fe. The
455 characterization of the metal composite material generated under Cu²⁺/Fe³⁺ irradiation and its
456 interaction with pharmaceuticals was outside the scope of this work but would deserve further
457 investigation.

458 4. Conclusions

459 This study investigated the fate of selected pharmaceuticals in Cu- and Fe-rich waters of
460 two Mediterranean intermittent rivers at lab- and field-scale. It is often challenging to transfer
461 kinetics from the laboratory to complex variable field conditions. However, in this study, only
462 photochemical processes due to Cu and Fe photoreactivity accounted for attenuation of investigated
463 pharmaceuticals and consequently, results on kphoto and katt agreed rather well. To go deeper in our
464 understanding of phototransformation mechanisms, TPs identification by LC-HRMS in irradiated
465 Fe³⁺/Cu²⁺ systems was suitable and highlighted specific and new transformation pathways including
466 reductive transformation pathways due to dehalogenation and hydrogenation reactions. Their
467 significance was confirmed by the detection of few of these TPs along the investigated river.

468 Acknowledgements
This research was supported by the EU PRIMA program through the research project INWAT –
Quality and management of intermittent river and groundwater in Mediterranean basins. The
authors thank the Platform of Non-Target Environmental Metabolomics (PONTEM) of the consortium
facilities Montpellier Alliance for Metabolomics and Metabolism Analysis (MAMMA). Ecole des Mines
d'Alès is thanked for the help in field sampling.

474

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<figure><img src="image_16.png" /><figcaption>Oxazepam</figcaption></figure>

<figure><img src="image_17.png" /><figcaption>Venlafaxine</figcaption></figure>

<figure><img src="image_18.png" /><figcaption>Fluconazole</figcaption></figure>

<figure><img src="image_19.png" /></figure>

<figure><img src="image_20.png" /><figcaption>5 TPs OXA</figcaption></figure>

<figure><img src="image_21.png" /><figcaption>9 TPs VEN</figcaption></figure>

<figure><img src="image_22.png" /></figure>

<figure><img src="image_23.png" /><figcaption>5 TPs FLU</figcaption></figure>

<figure><img src="image_24.png" /></figure>

Fe

Fe- and Cu-rich
Mediterranean
intermittent river

Photocatalytic redox transformation
dehalogenation, hydrogenation

10 TPs
CONFIRMED IN
RIVER SAMPLES

→ in-stream attenuation of pharmaceuticals