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# Reactive species-mediated phototransformation of triphenyl phosphate in cloud water: Kinetics, pathways and implications

Yi Wu, Feng Wu, Marcello Brigante, Angelica Bianco

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Yi Wu, Feng Wu, Marcello Brigante, Angelica Bianco. Reactive species-mediated phototransformation of triphenyl phosphate in cloud water: Kinetics, pathways and implications. Environmental Pollution, 2025, 383, pp.126717. 10.1016/j.envpol.2025.126717. hal-05229365

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# Reactive species-mediated phototransformation of triphenyl phosphate in cloud water: Kinetics, pathways and implications☆

Yi Wu<sup>a,b</sup>, Feng Wu<sup>c</sup>, Marcello Brigante<sup>a</sup>, Angelica Bianco<sup>b,*</sup> ◆

## ARTICLE INFO

### Keywords:

Atmospheric aqueous phase

Reactive species

Organophosphate esters

Indirect phototransformation

## ABSTRACT

Cloud water serves as crucial media for the transformation of organic compounds in the atmosphere, primarily through the oxidation of various reactive species. Triphenyl phosphate (TPHP) is one of the major organophosphate esters (OPEs) and is widely used in plasticizers and flame retardants, which has attracted extensive attention due to its toxicity and frequent occurrence in the atmosphere. However, its transformation pathway after partition into cloud water, especially mediated by reactive species, has not been reported. In this study, the kinetics, pathways, and atmospheric implications of TPHP transformation by hydroxyl radical (•OH), sulfate radical (SO₄•), and triplet excited states of organic matters (³C*) were investigated by using hydrogen peroxide (H₂O₂), persulfate (PDS), and 2-benzoylbenzoic acid (2-BBA) as precursors, respectively. Second-order rate constants of TPHP with these species were determined as k_TPHP,•OH = (7.20 ± 0.48) × 10⁹ M⁻¹ s⁻¹, k_TPHP,SO₄• = (5.99 ± 1.20) × 10⁹ M⁻¹ s⁻¹, k_TPHP,3C• = (1.73 ± 1.50) × 10⁹ M⁻¹ s⁻¹. Intermediate product analysis revealed the degradation pathway of TPHP, and short-chain carboxylic acids and phosphate ions were generated through the efficient oxidation of •OH. The contributions of •OH, SO₄•, and ³C* to the TPHP transformation in three typical cloud water environments (urban, remote, and marine) were evaluated by the pseudo-first-order rate. With the assessed TPHP lifetime of 94 min in urban areas, this study highlighted the dominant transformation process of organic contaminants in the atmospheric aqueous phase.

## 1. Introduction

Clouds, covering approximately 50 % of the Earth's surface, contribute 7 %-15 % of the atmospheric aqueous phase by volume (Lelieveld and Crutzen, 1991). They contain a large diversity of organic and inorganic compounds, metals, microorganisms, etc. Cloud droplets are tiny reactors that play an important role in atmospheric chemical composition and transformation, especially with solar irradiation (Bianco et al., 2020). Photogenerated reactive species drive the oxidation capacity of cloud water. Among them, hydroxyl radical (•OH) ($E_0$ = 1.8–2.7 V vs. NHE) has been considered the most efficient oxidant, which is produced mainly via the photolysis of hydrogen peroxide (H₂O₂), nitrates (NO₃), nitrites (NO₂) and metal complexes, as well as by direct uptake from the gas phase (Charbouillot et al., 2012; Wang et al., 2023). Steady-state concentrations of •OH in cloud water have been evaluated by model ranging from 10⁻¹⁵ to 10⁻¹² M (Herrmann

et al., 2010). Sulfate radical (SO₄•) ($E_0$ = 2.52–3.08 V vs. NHE), mainly produced through the oxidation chain of S(IV) is also significant, especially in polluted regions with higher SO₂ emissions (Otto et al., 2018; Tran et al., 2022). In addition, particular attention has also been focused on the triplet excited state (³C*) generated from the excitation of light-absorbing dissolved organics (Dalton et al., 2025; Felber et al., 2021; Kaur and Anastasio, 2018). They are primarily derived from biomass burning brown carbon and may compete with •OH in reactivity when ³C* is present in cloud water at concentrations between 10⁻¹⁵ and 10⁻¹³ M, especially for organics like phenols (Bretonniere et al., 2025; Ma et al., 2024). While previous works have typically focused on the transformation mechanism of organics by a single reactive species, •OH, SO₄•, and ³C* generally co-exist in the atmospheric aqueous phase, and their concentrations vary with the physicochemical composition of cloud water, which highly depends on the air mass origin (Vaïtilingom et al., 2013).

<footnote><sup>a</sup> CNRS, Institut de Chimie de Clermont-Ferrand, Université Clermont Auvergne, Clermont-Ferrand, 63000, France</footnote>

<footnote><sup>b</sup> CNRS, OPGC, Laboratoire de Météorologie Physique, UMR 6016, Université Clermont Auvergne, Aubière, 63178, France</footnote>

<footnote><sup>c</sup> School of Resource and Environmental Sciences, Wuhan University, Wuhan, 430079, China</footnote>

<footnote>* This paper has been recommended for acceptance by Charles Wong.</footnote>

<footnote>* Corresponding author. Laboratoire de Météorologie Physique (LaMP), 4 av. Blaise Pascal, Aubière, 63178, France.
E-mail address: angelica.bianco@uca.fr (A. Bianco).</footnote>
Organic compounds are one of the important sinks of reactive species in cloud water, while 70–90 % of the dissolved organics remain unidentified (Bianco et al., 2016; Sun et al., 2023). A growing concern in atmospheric chemistry is the occurrence of complex organic pollutants of both biogenic and anthropogenic origin, like phenolic compounds, pesticides, additives, etc., which have been widely detected in the atmosphere (Lebedev et al., 2018; Li et al., 2023). In particular, organophosphate esters (OPEs) have attracted attention due to their extensive production and application, potential bioaccumulation, and the concerning ecological and health risks (Fu et al., 2021; Ye et al., 2023). Global consumption of OPEs has surged from $1.9 \times 10^5$ tons in 2001 to $2.8 \times 10^6$ tons in 2018, resulting in the widespread occurrence in water, sediment, fish, and atmosphere, even being long-range atmospheric transport (LRAT) to remote and marine areas (e.g., the Arctic and Mediterranean and Black Seas) (Castro-Jiménez et al., 2014; Castro-Jiménez et al., 2016; Fu et al., 2023). Triphenyl phosphate (TPHP) is a typical aromatic OPE used in PVC plastics, polyurethane foams, hydraulic fluids, nail polish, and flame retardants, and easily released into the environment primarily due to physical rather than chemical bonding with materials (Fan et al., 2024; Kim and Kannan, 2018). TPHP represented one of the most prevalent OPEs observed in air samples, especially in electronic-waste-relevant areas (Bi et al., 2010; Yan et al., 2023). It was added to the candidate list of substances of very high concern (SVHC) by the European Chemicals Agency (ECHA) due to its endocrine disruption, neurotoxicity, reproductive toxicity, and carcinogenicity (Fu et al., 2021). It has been detected in the global atmosphere, both in gas phase: 170–580 $pg \cdot m^{-3}$ in Chicago, around 29.43 $pg \cdot m^{-3}$ in the Indo-China Peninsula, etc.; and in aerosol phase: 4–150 $pg \cdot m^{-3}$ in the North Sea, up to 1930 $pg \cdot m^{-3}$ in the Arctic Ocean, etc. (Castro-Jiménez et al., 2016; Peverly et al., 2015; Zhang et al., 2024). Although reports on TPHP concentrations in the atmospheric aqueous phase are limited, with vapor pressure of $1.6 \times 10^{-4}$ Pa, Henry's law constant of $3.3 \times 10^{-6}$ $atm \cdot mol^{-1}$, and $logK_{ow}$ of 4.59 (Table S1), our recent measurements in cloud water (e.g., $0.24 \pm 0.12 \mu g \cdot L^{-1}$ at Puy de Dôme, France) suggest that it is readily partitioned from the gas or aerosol phase into cloud water, where it can undergo transformation (Dominutti et al., 2022; Wang et al., 2020a). •OH has been considered the dominant atmospheric sink for TPHP. For example, the particle-bound TPHP was observed to degrade as a result of •OH exposure, and the derived second-order rate constant of TPHP with •OH was $(1.26 \pm 0.11) \times 10^9 \cdot M^{-1} \cdot s^{-1}$ (Liu et al., 2014). And gas phase transformation mechanism of TPHP initiated by •OH was modeled with estimated rate constant and lifetime (Yu et al., 2016). Nevertheless, the transformation mechanism of TPHP in the atmospheric aqueous phase, especially mediated by multiple reactive species ($\cdot OH$, $SO_4^{2-}$, and $^3C^*$) remains unclear.

To address the gap, this work employed hydrogen peroxide ($H_2O_2$), peroxydisulfate (PDS), and 2-benzoylbenzoic acid (2-BBA) as precursors to generate •OH, $SO_4^{2-}$, and $^3C^*$ respectively under simulated solar irradiation. We investigated the oxidation reactivity of atmospheric aqueous phase and the fate of emerging contaminants, focusing on TPHP as a case study, by i) determining the second-order reaction rate constants of •OH, $SO_4^{2-}$, and $^3C^*$ with TPHP by competitive kinetics and laser flash photolysis; ii) identifying intermediate products of TPHP oxidation and clarifying degradation pathways; iii) evaluating the contribution of •OH, $SO_4^{2-}$, and $^3C^*$ to the transformation of TPHP in three typical cloud water (urban, remote, and marine) to provide insights into its atmospheric photochemical fate.

## 2. Materials and methods

### 2.1. Chemicals and reagents

All chemicals were directly used without further purification. Triphenyl phosphate (TPHP, ≥99 %), hydrogen peroxide solution ($H_2O_2$, 30 % w/w), sodium persulfate (PDS, ≥98 %), terephthalic acid (TA,

≥98 %), benzoic acid (BA, ≥99 %), and acetonitrile (ACN) of HPLC grade were purchased from Sigma-Aldrich. 2-benzoylbenzoic acid (2-BBA, >98 %) was obtained from Interchim, France. Formic acid (HCOOH, >98 %) was obtained from BDH Prolabo, VWR. Aqueous solutions were prepared in Milli-Q water (>18 MΩ cm) and stored in the dark at 277 K for further use. Since the solubility of TPHP is limited, the supernatant was freshly obtained by filtering (5.0 µm pore size, hydrophilic PVDF, Millipore) the well-stirred and dissolved TPHP solution, then analyzed and quantified. The concentration of $H_2O_2$ stock solution was determined with a molar absorption coefficient of $38.1 \pm 1.4 \cdot M^{-1} \cdot cm^{-1}$ at 240 nm (Miller and Kester, 1988).

### 2.2. Irradiation procedure

Irradiation experiments were performed using the water-jacketed reactor (45 mL) equipped with magnetic stirring and temperature maintaining (293 K) system under a 1000 W Xenon lamp simulating solar light with a Pyrex filter removing wavelength lower than 285 nm. The emission spectrum of the light reaching reaction solution was recorded with a CCD spectrophotometer (Ocean Optics USD, 2000) and was comparable in wavelength and energy to natural sunlight at Puy de Dôme observatory (1465 m a.s.l.) in clear and cloudy sky conditions. The UV–Vis absorption spectra of 5 µM TPHP, 0.5 mM $H_2O_2$, 2-BBA, and PDS aqueous solutions were shown in Fig. S1. The reaction solutions containing desired concentrations of TPHP and $H_2O_2$/2-BBA/PDS were completely mixed just before light exposure and 0.5 mL aliquot of the solution was withdrawn at the defined time for the analysis of TPHP concentrations.

### 2.3. Analytical methods

The residual concentrations of TPHP were measured by High Performance Liquid Chromatography equipped with a diode-array detector (HPLC, 1260 Infinity II, Agilent) and an eclipse XDB-C18 column (4.6 × 150 mm, 5 µm, Agilent). The gradient elution at a flow rate of 1 mL min⁻¹ was as follows: start with 40 % ACN and 60 % acidified water (0.1 % HCOOH) for 3 min, followed by a linear gradient to 90 % ACN within 12.5 min, maintained for 2 min, and with a linear gradient back to 40 % ACN within 0.5 min, then maintained for 3 min. The retention time of TPHP was 13 min with column temperature at 308 K and detection wavelength at 260 nm.

TPHP transformation intermediate products were analyzed by Ultra-High Performance Liquid Chromatography-Mass Spectrometry (Ultimate 3000 RSLC, Thermo Scientific) equipped with ESI with positive and negative ionization mode over *m/z* = 50–750. The injection volume was 5 µL. Separation was performed on a C18 column (4.6 × 250 mm, 1.7 µm, Kinetex) at a flow rate of 0.45 mL min⁻¹ with ACN and acidified water (0.1 % HCOOH) as gradient elution solvents.

Short-chain carboxylic acids and phosphate ions ($PO_4^{3-}$) were identified by Ion Chromatography-Mass Spectrometry (IC-MS, Thermo Scientific) equipped with Dionex ICS-6000 system, single quadrupole mass spectrometer, and Dionex IonPac AS11-HC-4 µm column (2 × 250 mm). The gradient elution was carried out using KOH solutions of different concentrations at a flow rate of 0.36 mL min⁻¹. The column temperature was set at 313 K and the injection volume was 40 µL.

### 2.4. Competitive kinetics method

The second-order rate constants between TPHP and reactive species (RS) ($k_{TPHP,RS}$) were determined by competitive kinetics experiments (Aschmann et al., 2006; Passananti et al., 2014). •OH and $SO_4^{2-}$ can react with TPHP and reference compounds (TA or BA) following the second-order rate law, where [TPHP] and [Ref] represent concentrations, respectively. $k_{TPHP,RS}$ ($k_{TPHP,\cdot OH}$ and $k_{TPHP,SO_4^{2-}}$) can be calculated with the reported constants ($k_{Ref,RS}$) according to Eq. (1). Reaction
solutions containing 5 µM TPHP, 10 µM TA or 100 µM BA, and 500 µM H₂O₂ or PDS were irradiated in the system described in Section 2.2, and samples were withdrawn at selected time intervals for HPLC analysis of TPHP and TA or BA.

$$ \ln \frac{[TPHP]_t}{[TPHP]_0} = \frac{k_{TPHP,RS}}{k_{Ref,RS}} \times \ln \frac{[Ref]_t}{[Ref]_0} \qquad (1) $$

## 2.5. Laser flash photolysis (LFP)

The excited state of 2-BBA (³²2-BBA*) was analyzed by transient absorption spectroscopy using LFP system (Quanta Ray GCR 130-01 Nd: YAG) operated with the fourth harmonic (λexc = 266 nm) (Fig. S2). The single pulses were 9 ns in duration with a typical energy of 50 mJ/pulse. The second-order rate constant between TPHP and ³²2-BBA* (kTPHP,³²2-BBA*) was derived by fitting the first-order decay constant of ³²2-BBA* (k³²2-BBA*) monitored at 320 nm as a function of the initial TPHP concentration (Eq. (2)), while k³²2-BBA* was determined by fitting the absorbance versus time with exponential equations (Berruti et al., 2022). Stock solutions of TPHP and 2-BBA were completely mixed just before each experiment to obtain the desired concentrations in the air-equilibrated solution of ACN and Milli-Q water (1:1, v/v) at room temperature (295 ± 2) K. The details of the LFP agree with our previous studies and are presented in Text S1 (De Laurentiis et al., 2013).

$$ k_{32-BBA^*} = k_{TPHP,32-BBA^*} [TPHP] + k_{32-BBA^*,0} \qquad (2) $$

# 3. Results and discussion

## 3.1. Phototransformation kinetics of TPHP

According to previous studies, H₂O₂ is the main source of •OH in cloud water and was used as a precursor to generate •OH under simulated solar irradiation to investigate the •OH-mediated photolysis of TPHP (Bianco et al., 2015). With the initial H₂O₂ concentration increased from 0.1 to 0.5 mM, the degradation efficiency of TPHP increased from 49.7 % to 100 % after 120 min reaction (Fig. 1a), and no significant removal of TPHP was observed by direct photolysis or reaction with H₂O₂ in dark (Fig. S3). The time evolution of TPHP followed a pseudo-first-order kinetics, and the apparent rate constants (kobs) for the first 30 min of reactions were calculated with Eq. (3) (Fig. S4), which were plotted against the initial H₂O₂ concentration in Fig. 1d. As the concentration of H₂O₂ increased from 0.1 to 1 mM, kobs increased linearly from 0.0068 to 0.0498 min⁻¹, shown as the dashed line. However, it was 0.0695 min⁻¹ with 2 mM H₂O₂ due to that the excessive H₂O₂ can consume •OH and the produced HO₂⁻ is less reactive toward TPHP (Buxton et al., 1988). Meanwhile, 0.5 mM TA was introduced to scavenge •OH in the irradiation system (Text S2) and •OH production rate (R₀H) was quantified as 1.56 × 10⁻⁸ M s⁻¹ (Fig. S5), while it was measured as ranging from 2 × 10⁻¹¹ to 4 × 10⁻¹⁰ M s⁻¹ in natural cloud water (Bianco et al., 2015).

<figure><img src="image_3.png" /><figcaption>Fig. 1. Effect of H₂O₂ (a), PDS (b), and 2-BBA (c) on TPHP transformation under simulated solar irradiation. Trends of apparent rate constants (kobs) vs. initial precursor concentrations (d). Initial conditions: [TPHP] = 5 µM, [H₂O₂] = 0.1–2 mM, [PDS] = 0.05–0.5 mM, [2-BBA] = 0.1–0.5 mM.</figcaption></figure>
$$ \ln \left[ \frac{\text{TPHP}}{\text{TPHP}} \right]_0 = -k_{\text{obs}} \times t \qquad (3) $$

$SO_4^2-$ in the atmospheric aqueous phase is derived primarily from sulfur dioxide autoxidation chain and minor sources of sulfate or bisulfate ions activation by $\cdot$OH, $NO_3^-$, iron, etc (Otto et al., 2018; Tran et al., 2022). Recently, some new formation mechanisms, such as the photolysis of sulfate aerosols and the interface reactions of ammonium sulfate aerosol particles, have also been revealed (Cope et al., 2022; Kong et al., 2021). In this work, $SO_4^2-$ was generated from PDS under simulated solar irradiation at natural weak acid pH (5.0–6.0), with negligible production of $\cdot$OH (Berruti et al., 2022; Otto et al., 2018). Results showed that PDS significantly accelerated TPHP photo-transformation efficiency, which was completely removed just after 20 min with 0.5 mM PDS (Fig. 1b). $k_{\text{obs}}$ according to the pseudo-first-order kinetics (Fig. S6) grew from 0.0235 to 0.1730 min$^{-1}$ over the PDS concentration range of 0.05–0.5 mM (Fig. 1d). In contrast to $\cdot$OH, concentrated PDS did not slow down the linear increase of $k_{\text{obs}}$, since the quenching effect of PDS on $SO_4^2-$ is limited ($K_{S_2O_8^{2-} \cdot SO_4^{2-}} = 6.0 \times 10^5 \text{ M}^{-1} \text{s}^{-1}$) (Lin et al., 2024).

Brown carbon, which is mainly produced from biomass burning, commonly includes compounds with aromatic carbonyls, that can be photosensitized to excited states in the atmosphere (Felber et al., 2021; Kaur and Anastasio, 2018; Li et al., 2023). BBA is a well-documented model aromatic ketone photosensitizer that has a triplet reduction potential close to that of natural chromophoric dissolved organic matter (CDOM) (Bretonniere et al., 2025; Dalton et al., 2025; Petersen-Sonn et al., 2024). At natural pH (3.0–3.5), 2-BBA is approximately half ionized (Fig. S7) and the phototransformation of TPHP increased with 2-BBA concentration (Fig. 1c), especially during the initial 15 min (Fig. S8). However, it was quite slow at neutral pH (Fig. S9), related to the species distribution of the ground and triplet excited states of 2-BBA. This result is consistent with our recent work on the reactivity of $^3$2-BBA* with carboxylic acids and amino acids using LFP (Bretonniere et al., 2025).

$$ 2-\text{BBA} \xrightarrow{h\nu_{\text{ISC}}} ^{h\nu}_{^{3}}2-\text{BBA}^* \qquad \text{R1} $$

$$ ^{3}2-\text{BBA}^* + \text{TPHP} \rightarrow 2-\text{BBA}^{\cdot\cdot} + \text{TPHP}^{\cdot\cdot} \rightarrow \text{Products} \qquad \text{R2} $$

For the photosensitized process, the ground state of 2-BBA is excited into triplet state ($^32$-BBA*) through intersystem crossing (ISC) (R1), which could further 1) directly oxidize TPHP via electron transfer (R2); 2) undergo energy transfer with molecular oxygen to form singlet oxygen ($^1O_2$); 3) deactivate due to contact with water and itself (Li et al., 2017b). Notably, $^1O_2$ is also a key reactive species in the natural environment, whose steady-state concentrations in fog water are around 3 times that of $^3CDOM$* and 100 times that of $\cdot$OH (Kaur and Anastasio, 2017; Kaur and Anastasio, 2018). However, $^1O_2$ is a moderately active electrophile and generally exhibits selectivity, favoring electron-rich organics (Tratnyek and Hoigne, 1991; Wang et al., 2024). In this work, Rose Bengal (RB) was used as a photosensitizer to specifically produce $^1O_2$ and investigated its effect on TPHP (Text S3). With more than 40 % of RB decomposition after 7 h of photoexcitation, TPHP was not degraded (Fig. S10). However, the reaction system showed efficiency toward FFA, a common probe compound for $^1O_2$, indicating that $^1O_2$ ($E(^1O_2/O_2^\cdot) = 0.81 \text{ V vs. NHE}$) hardly contributed to the transformation of TPHP (Dong et al., 2021). In addition, $O_2^\cdot$ ($E(O_2^\cdot/H_2O_2) = 0.91 \text{ V vs. NHE}$) and $H_2O_2$ that might be produced by the further reaction of $^32$-BBA* in the presence of $O_2$ have also been proved unlikely to directly oxidize TPHP; thus the phototransformation of TPHP could be considered as the role of $^32$-BBA*. The photolysis rates of 2-BBA (0.1–0.5 mM) were identical likely due to significant UV–Vis absorption of 2-BBA that may result in a light screening effect (Fig. S11). Notably, the transformation rate of TPHP slowed after the initial 20 min and the kinetics were similar at different 2-BBA concentrations, possibly because the self-quenching reaction of $^32$-BBA* ($k_{^32-\text{BBA}^* \cdot 2-\text{BBA}} = 2.22 \times$

$10^9 \text{ M}^{-1} \text{s}^{-1}$) gradually dominated with the consumption of TPHP (Bretonniere et al., 2025).

## 3.2. Second-order rate constants

As mentioned above, reactive species present in cloud water, including $\cdot$OH, $SO_4^2-$, $^3C^*$, all showed effective transformation for TPHP. Previous works have focused on the reaction of $\cdot$OH with TPHP in the gas phase ($k = 1.6 \times 10^{-12} \text{ cm}^3 \text{ molecule}^{-1} \text{ s}^{-1}$) and at the surface of aerosol particles ($k = (2.1 \pm 0.2) \times 10^{-12} \text{ cm}^3 \text{ molecule}^{-1} \text{ s}^{-1}$) (Liu et al., 2014; Yu et al., 2016). The rate constant in atmospheric aqueous phase and the reactivity with $SO_4^2-$ and $^3C^*$ have not been reported to our knowledge. Competition kinetic is an effective approach for second-order rate constant determination (Aschmann et al., 2006; Pasananti et al., 2014). Fig. S12 showed the removal efficiency of TPHP and TA with 500 µM $H_2O_2$ under irradiation. With Eq. (1) and the reported rate constant ($k_{TA,\cdot\text{OH}} = 4.0 \times 10^9 \text{ M}^{-1} \text{ s}^{-1}$), the slope of the linear fitting in Fig. 2a gave $k_{\text{TPHP},\cdot\text{OH}} = (7.20 \pm 0.48) \times 10^9 \text{ M}^{-1} \text{ s}^{-1}$ (Charbouillot et al., 2011). While it varied from the reported $k_{\text{TPHP},\cdot\text{OH}}$ in UV/$O_3$ oxidative digestion system of (7.1 ± 1.0) × $10^8 \text{ M}^{-1} \text{ s}^{-1}$ and the determined bimolecular reaction rate constants of other OPEs, including TCEP, TCP/IP, TDCPP with $\cdot$OH, ranging from 2.5–4.0 × $10^{10} \text{ M}^{-1} \text{ s}^{-1}$ (Tang et al., 2019; Zhu et al., 2024).

$$ k_{\text{C}+\text{OH},\text{gas}}[\text{cm}^3 \text{ molecule}^{-1} \text{ s}^{-1}] = 1.83 \times 10^{-21} k_{\text{C}+\text{OH},\text{aqueous}} [\text{M}^{-1} \text{ s}^{-1}] \qquad (4) \\ R^2 = 0.77 \quad p < 0.05 $$

Since TPHP has both aromatic and OPE properties, a linear regression relationship was developed (Fig. S13) between the rate constants of compounds similarly structured to TPHP with $\cdot$OH in the atmospheric gas ($k_{\text{C}+\cdot\text{OH},\text{gas}}$) and aqueous phase ($k_{\text{C}+\cdot\text{OH},\text{aqueous}}$) (Eq. (4)) (Vione, 2020). Where $k_{\text{C}+\cdot\text{OH},\text{aqueous}}$ was obtained from the reported measurements, and $k_{\text{C}+\cdot\text{OH},\text{gas}}$ was calculated using the shareware software EPI Suite by US-EPA based on Quantitative Structure-Activity Relationships (QSAR) assessment, listed in Table S2 (Li et al., 2023; USEPA, 2019). $k_{\text{TPHP},\cdot\text{OH},\text{gas}}$ ($1.08 \times 10^{-11} \text{ cm}^3 \text{ molecule}^{-1} \text{ s}^{-1}$) was also estimated by EPI Suite and applied to Eq. (4), and the obtained $k_{\text{TPHP},\cdot\text{OH},\text{aqueous}$ ($5.92 \times 10^9 \text{ M}^{-1} \text{ s}^{-1}$) was within one order of magnitude of the experimental value in this work ($k_{\text{TPHP},\cdot\text{OH}} = (7.20 \pm 0.48) \times 10^9 \text{ M}^{-1} \text{ s}^{-1}$).

For $SO_4^2-$, $k_{\text{TPHP},\text{SO}_4^{\cdot\cdot}-}$ was calculated following the similar protocol adopted for $\cdot$OH. In addition to TA ($k_{\text{TA,SO}_4^{\cdot\cdot}-} = 1.7 \times 10^8 \text{ M}^{-1} \text{ s}^{-1}$), BA ($k_{\text{BA,SO}_4^{\cdot\cdot}-} = 1.2 \times 10^9 \text{ M}^{-1} \text{ s}^{-1}$) was also used as the reference compound to trap $SO_4^2-$ in $h\nu$-PDS-TPHP system (Fig. S14) (Guan et al., 2011; Yan et al., 2024). According to the slopes of linear fit in Fig. 2b, although the rate constants of TA and BA as probes differ by nearly 10-fold, the obtained $k_{\text{TPHP,SO}_4^{\cdot\cdot}-}$ are close ((7.19 ± 0.83) × $10^9 \text{ M}^{-1} \text{ s}^{-1}$ and (4.79 ± 0.40) × $10^9 \text{ M}^{-1} \text{ s}^{-1}$). Therefore, the average value ($k_{\text{TPHP,SO}_4^{\cdot\cdot}-} = (5.99 \pm 1.20) \times 10^9 \text{ M}^{-1} \text{ s}^{-1}$) is considered in the following. This value agrees with the reported range of bimolecular rate constants ($10^8-10^9 \text{ M}^{-1} \text{ s}^{-1}$) between $SO_4^2-$ and electron-rich molecules (e.g., aromatic compounds) via a single electron transfer mechanism (Wojnárovits and Takács, 2019). The rate constant between TPHP and $SO_4^2-$ is comparable to that of $\cdot$OH, even if slightly lower. However, with the same amount of precursor, like 0.5 mM PDS or $H_2O_2$, the phototransformation efficiency of TPHP in the $h\nu$-PDS system was significantly higher than that of the $h\nu$-$H_2O_2$ system (Fig. 1), which is probably related to the photolysis quantum yield and longer half-life of $SO_4^2-$ (30–40 µs) (Berruti et al., 2022; Chen et al., 2020). Yan et al. applied LFP to determine the quantum yields of PDS and $H_2O_2$ for the production of $SO_4^2-$ and $\cdot$OH, which were 1.46 and 1.10, respectively. Density functional theory (DFT) calculations further indicated that PDS is more prone to bond cleavage and radical formation than $H_2O_2$ (Yan et al., 2024).

Time-resolved laser flash photolysis (LFP) has been used for transient species kinetic investigations, especially widely for the transformation of organics triggered by radicals or excited triplet states, thanks to its
<figure><img src="image_4.png" /><figcaption>**Fig. 2.** The linear fit of TPHP and TA transformation rate in the *hv*-H₂O₂ system (**a**). The linear fit of TPHP and TA or BA transformation rate in the *hv*-PDS system (**b**). Transient absorption spectra of 2-BBA upon LFP (266 nm, 50 mJ) with and without 600 µM TPHP (**c**). The first order decay constant of ³²BBA* (k₃₂-BBA*) as a function of initial TPHP concentration (**d**). Solid line showed linear fitting, and dashed line showed 95 % confidence interval.</figcaption></figure>

higher temporal resolution, faster response speed, and lower detection limit (Berruti et al., 2022; Yan et al., 2024). ³²BBA* was produced *in situ* upon excitation of an aqueous solution containing 0.1 mM 2-BBA at 266 nm. The maximum absorption of ³²BBA* at natural pH lies at 320 nm based on our previous study (Bretonniere et al., 2025). Accordingly, the signal was monitored at 320 nm (Fig. 2c), permitting the determination of triplet state decay with a pseudo-first-order constant (k₃₂-BBA*). ⁷·⁸⁶ × 10⁶ s⁻¹ and a lifetime of 127 ns in our experimental conditions. The decay of ³²BBA* was accelerated with the increasing initial concentrations of TPHP ranging of 0.1–0.6 mM (Fig. 2d), and kₗTPHP,³₂-BBA* was determined as (1.73 ± 1.50) × 10⁹ M⁻¹ s⁻¹ from the slope, which is comparable to that of *OH and SO₄* (kₗTPHP,•OH > kₗTPHP,SO₄• > kₗTPHP,³₂-BBA*). (Table S3).

### 3.3. Phototransformation products and pathways

Transformation products of TPHP were identified by HPLC-MS analysis which was operated at both positive and negative ionization modes ([(M+H)]⁺ and [M-H]⁻). The reaction time was: 30, 60, 120 min in *hv*-H₂O₂ system; 5, 20 min in *hv*-PDS system; 15, 120 min in *hv*-2-BBA system. The total ion chromatograms (TICs) of the control sample, which only contained 5 µM TPHP and without oxidants, showed no detectable

products. In the *hv*-H₂O₂ system, in addition to TPHP (retention time = 5.71 min), 6 molecular formula compounds with 10 separate retention time were detected (Table S4), suggesting the presence of isomers with OH-substituted at different sites of the phenyl ring. Generally, *OH tends to react with organics, especially aromatics, via direct addition, H-atom abstraction, and electron transfer mechanisms (Song et al., 2019; Zhang et al., 2023). The mono-hydroxyl adducts (P1, P2) observed at retention time of 5.12 and 5.01 min (m/z 342.07) were likely para- and meta-substituted derivatives (Yu et al., 2016). The di-hydroxyl adducts (P3, P4, P5) appeared at 4.72, 4.53, and 4.43 min (m/z 358.06), with OH could distribute on one or separate phenyl ring, while no tri-hydroxyl adduct was detected. It is understood that the hydroxyl adducts were more susceptible oxidized and cleaved than to further undergo hydroxyl substitution (Dey et al., 2019). Diphenyl phosphate (DPHP, P6), resulting from the P-O bond cleavage when P-atom was attacked by *OH, was observed at 4.23 min (m/z 250.04). Meanwhile, the mono (P7, P8, m/z 266.03) and di (P9, m/z 282.03) hydroxyl adducts of DPHP have also been found at 3.49, 3.14, and 3.12 min, respectively, which could also be generated from TPHP-OH (P1~P5). Moreover, phenylphosphate (P10), the product from further cleavage of the P-O bond and removal of a phenyl group, occurred at 1.08 min (m/z 174.01). Based on the detected molecular formula and presumed structure of products, the transformation pathways of TPHP were proposed in Fig. 3.
<figure><img src="image_5.png" /><figcaption>**Fig. 3.** Proposed transformation pathways of TPHP by •OH, $SO_4^{2-}$, and $^32$-BBA*. Initial conditions: [TPHP] = 5 µM, [H$_2$O$_2$] = 0.5 mM, [PDS] = 0.5 mM, [2-BBA] = 0.5 mM under irradiation.</figcaption></figure>

Remarkably, with TPHP-OH and DPHP being detected in both *hv*-PDS and *hv*-2-BBA systems (Tables S5 and S6), revealing that $SO_4^{2-}$ and $^32$-BBA* could contribute to the transformation of TPHP by electrophilic attack on aryl-C atom and bridging-O atom. $SO_4^{2-}$ mainly reacts with chemicals through electron transfer mechanism, thus it could extract the electron from the phenyl group, generating TPHP cation radical

(TPHP*+), which rapidly reacted with water to form TPHP-OH (Chen et al., 2020; Song et al., 2019). Meanwhile, the di-hydroxyl adducts, which were overlooked in previous work, have been first detected. $SO_4^{2-}$ could also attack the O-atom, leading to the P-O bond cleavage to form DPHP. DPHP-OH was produced through the further oxidation of TPHP-OH and DPHP following the same pattern (Dong et al., 2021). It

<figure><img src="image_6.png" /><figcaption>**Fig. 4.** Short-chain carboxylic acids and phosphate ions production in the *hv*-H$_2$O$_2$ system (a). Removal rate of TPHP (in 30 min) and formation rate of short-chain carboxylic acids (in 120 min) (b). Initial conditions: [TPHP] = 5 µM, [H$_2$O$_2$] = 0.5 mM under irradiation.</figcaption></figure>
has been demonstrated that triplet excited states (³C*) can oxidize organic compounds through electron transfer and hydrogen-atom abstraction mechanisms (Foote, 1968; Li et al., 2017b). Accordingly, it could be explained that TPHP was oxidized by ³2-BBA*, giving radicals (TPHP•⁺ and 2-BBA•), which then react with water to form TPHP-OH (P1~P5) (Dalton et al., 2025). Similarly to SO₄²⁻, ³2-BBA* could also lead to the cleavage of the P-O bond by electrophilic attack and giving DPHP (P6), though the products of further hydroxylation have not been detected.

In addition, short-chain carboxylic acids and phosphate (PO₄³⁻) were generated in the hν-H₂O₂ system (Fig. 4a, Table S7). Notably, with the efficient removal of TPHP, the formation rates (r) of short-chain carboxylic acids followed r (C1) > r (C2) > r (C3) > r (C4), and PO₄³⁻ started to be detected after 30 min (Fig. 4b). The results suggested that at the initial stage of the reaction, the oxidation of TPHP by •OH probably caused the cleavage of phenyl group, releasing C1 compound (formic acid) and simultaneously generating some phosphate ester intermediates (Chen et al., 2024; Li et al., 2023). These compounds could be further oxidized and rapidly dephosphorylated after 30 min, consistent with the results that phosphorus-free aromatic compounds were undetectable by HPLC-MS. Regarding SO₄²⁻, while TPHP was completely photo-transformed in 30 min in the presence of 0.3 mM PDS (Fig. 1c), the production of short-chain carboxylic acids was limited and PO₄³⁻ was undetectable after 2 h (Fig. S15). It suggested that SO₄²⁻ (E₀ = 2.52–3.08 V vs. NHE) exhibited selectivity as an electrophilic radical, thus the role of SO₄²⁻ on the oxidation and fate of emerging contaminant in the atmospheric aqueous phase required further investigation, as does the triplet excited states (Titaley et al., 2019; Tran et al., 2022).

### 3.4. Atmospheric implications

The pH range in cloud water tends to the neutral in the last decade, as anthropogenic emissions are controlled and the medium is generally diluted by high liquid water content (Pye et al., 2020). The measured pH values, between 4 and 7, may lead to a slightly lower reactivity of triplet excited states (³C*), which are more reactive in acidic conditions (Smith et al., 2014). Nevertheless, fog droplets and deliquescent aerosols are more acidic and concentrated and the photogenerated ³C* may play an important role in the degradation of organics (Li et al., 2023).

•OH, SO₄²⁻, and ³C* present in the atmospheric aqueous phase can efficiently degrade TPHP (~10⁹ M⁻¹ s⁻¹), though the oxidation mechanism and intensity may differ, leading to complete mineralization only in the case of •OH. Considering that the pH of cloud water ranges widely depending on its source and chemical composition, in this work, k_TPHP,RS determined at natural pH (6.0–7.0 for •OH, 5.0–6.0 for SO₄²⁻, and 3.0–3.5 for ³C*) were used to preliminarily estimate the transformation of TPHP in the oxidative atmospheric aqueous phase (dashed line in Fig. 5a). Meanwhile, the abundance of reactive species in cloud water formed by different air mass origin (urban, remote, marine) also varies (cube in Fig. 5a). Therefore, the first-order transformation rate of TPHP was estimated based on the steady-state concentrations of •OH and SO₄²⁻ calculated by CAPRAM model and that of measured ³C* (Herrmann et al., 2010; Kaur and Anastasio, 2018), listed in Table S8. The assessed contribution of •OH, SO₄²⁻, and ³C* to the transformation of TPHP in three typical origins of cloud water was shown in Fig. 5b. Notably, ³C* contributes nearly 50 % in urban areas, which means that in polluted areas, with the emission and accumulation of various phenolic compounds, and a slightly decreased pH, ³C* would play a more significant role (Wang et al., 2020b). Conversely, in clean remote or marine areas, •OH could play a dominant role in the transformation of TPHP due to the lower concentration of light-absorbing organics, which may lead to a lower level of photogenerated ³C* and a higher level of •OH because of

<figure><img src="image_7.png" /></figure>

<figure><img src="image_8.png" /><figcaption>Fig. 5. Second-order rate constants of TPHP with •OH, SO₄²⁻, ³C* (dashed line) and the range of steady-state concentrations of •OH, SO₄²⁻, ³C* as reported (cube) (Herrmann et al., 2010; Kaur and Anastasio, 2018) (a). Assessment of the contribution of •OH, SO₄²⁻, and ³C* to TPHP transformation in typical origins of cloud water (b).</figcaption></figure>

the limited organics acting as radical scavengers (Herckes et al., 2013). More field measurements of ³C* in various environments are needed to better assess the implications.

$$ \tau_{\text{TPHP}} = \frac{1}{k_{\text{TPHP,OH}}[\cdot\text{OH}] + k_{\text{TPHP,SO4}}[\text{SO4}] + k_{\text{TPHP,3C*}}[\text{3C*}]} \quad (5) $$

Furthermore, the lifetime of TPHP (τ_TPHP) was estimated to be 94 min based on Eq. (5) by combining the steady-state concentrations of •OH, SO₄²⁻, and ³C* in urban areas. This is shorter than the previously reported lifetimes at the surface of aerosol particles (τ_TPHP, particle = 5.6 days) and in the gas phase (τ_TPHP, gas = 7.6 days) (Table 1) (Liu et al., 2014; Yu et al., 2016). This gap could be due to the fact that SO₄²⁻ and ³C* were not taken into account for the calculations as reported, but could also highlight the enhanced reactivity of TPHP in the aqueous phase compared to the particle surface and gas phase. TPHP, which is primarily emitted in urban areas, could be transported over long distances
<table><caption>**Table 1**
Estimated lifetime of TPHP ($\tau_{TPHP}$) in different atmospheric phases.</caption><thead><tr><th>Phases</th><th>Method for k<sup>*</sup><sub>TPHP,OH</sub></th><th>k<sup>*</sup><sub>TPHP,OH</sub></th><th>[*OH]</th><th>Other Reactive Species</th><th>Area</th><th>τ<sub>TPHP</sub></th><th>References</th></tr></thead><tbody><tr><td>Aerosol surface</td><td>Experiment</td><td>2.1 × 10<sup>-12</sup> cm<sup>3</sup> molecule<sup>-1</sup> s<sup>-1</sup></td><td>1.0 × 10<sup>6</sup> molecules cm<sup>-3</sup></td><td>\</td><td></td><td>5.6 days</td><td>Liu et al. (2014)</td></tr><tr><td>Gas</td><td>Computation (DFT)</td><td>1.6 × 10<sup>-12</sup> cm<sup>3</sup> molecule<sup>-1</sup> s<sup>-1</sup></td><td>9.7 × 10<sup>5</sup> molecules cm<sup>-3</sup></td><td>\</td><td></td><td>7.6 days</td><td>Yu et al. (2016)</td></tr><tr><td>Gas</td><td>QSAR</td><td>1.1 × 10<sup>-11</sup> cm<sup>3</sup> molecule<sup>-1</sup> s<sup>-1</sup></td><td>1.5 × 10<sup>6</sup> molecules cm<sup>-3</sup></td><td>\</td><td></td><td>0.7 days</td><td>EPI Suite</td></tr><tr><td>Aqueous</td><td>Experiment</td><td>7.20 × 10<sup>9</sup> M<sup>-1</sup> s<sup>-1</sup></td><td>3.5 × 10<sup>-15</sup> M</td><td>SO<sub>4</sub><sup>2-</sup> and <sup>3</sup>C*</td><td>Urban<br>Remote<br>Marine</td><td>94 min<br>43 min<br>1.2 min</td><td>This work</td></tr></tbody></table>

at the particle surface and in the gas phase to remote and marine areas, where it is rapidly oxidized by *OH to short-chain carboxylic acids and PO₄³⁻ when it dissolves into the aqueous phase. The dissolution of TPHP in cloud water may be enhanced by the presence of lipophilic compounds at the air-water interface, with an oversaturation compared to the prediction with Henry's law constant, as already observed for other aromatic compounds (Dominutti et al., 2022; van Pinxteren et al., 2005; Wang et al., 2020a). It is worth noting that some emerging contaminants detected in the atmosphere may have their abundance and pollution overestimated due to their rapid decay and short lifetime in the aqueous phase.

In addition, the complex inorganic and organic components in cloud water could potentially accelerate or inhibit the transformation of contaminants (Deguillaume et al., 2014; Gioda et al., 2011). For example, highly marine cloud water is rich in Cl⁻ (~446 µM on average, but up to 841 µM), which can scavenge *OH (k = 4.3 × 10⁹ M⁻¹ s⁻¹) and SO₄²⁻ (k = 1.3–3.1 × 10⁸ M⁻¹ s⁻¹), and selectively react with <sup>3</sup>C* (Chen et al., 2020). NO₃⁻ produced by the photolysis of NO₃⁻ (~910 µM on average, but up to 8.7 mM) in polluted cloud water may also act as an oxidant (Li et al., 2017a). Another important factor of photochemical reactions that should not be overlooked is pH. Although TPHP was investigated here as a case study, whose structure with three phenyl rings results in weak ionization in water, the effect of pH on its reactivity would be minor (Song et al., 2019). But when we broaden our focus to more emerging contaminants, the effect of pH could be significant. It has been confirmed that for cis-pinonic acid, within the temperature range of 283–318 K, the ratio of k<sub>•OH</sub> at pH > 8 and pH < 2 stabilized at 0.92–1.19, and that of k<sub>SO₄•-</sub> was 0.95–1.36, suggesting that k<sub>•OH</sub> and k<sub>SO₄•-</sub> depended weakly on pH (Otto et al., 2018). For <sup>3</sup>C*, previous work using 3,4-dimethoxybenzaldehyde (DMB) as a model compound found that pH strongly affected the transformation rate of phenol, increasing nearly 20-fold from pH 6 to pH 2 (Smith et al., 2014). It is attributed to the higher content of protonated triplet states in acidic solutions, which are more reactive. Since the steady-state concentrations of *OH, SO₄²⁻, and <sup>3</sup>C* used in this work were derived from models that have already considered different chemical scenarios, the assessment is relatively accurate but experimental measurements of the reactive species in natural samples collected in contrasting environment should be further investigated.

## 4. Conclusions

To the best of our knowledge, this study provided the first insights into the transformation of emerging contaminants mediated by reactive species in cloud water, with a particular focus on TPHP, a representative aryl OPE detected in atmospheric air, aerosol, rain, and cloud. Based on the determined second-order rate constants (k<sub>TPHP,•OH</sub> > k<sub>TPHP, SO₄•-</sub> > k<sub>TPHP,3,2-BBA</sub>), the triplet excited state is estimated to contribute nearly 50 % to TPHP transformation in polluted urban areas with higher abundance of light-absorbing organics. In contrast, *OH would dominate following long-range atmospheric transport to remote or marine regions. TPHP-OH and DPHP were identified as the major transformation products, where the extensive oxidation by *OH could further generate short-chain carboxylic acids and phosphate ions, suggesting the potential role that photochemical transformation of organophosphate esters

could play in the phosphorus cycle. Moreover, this work addresses the research gap previously limited to the gas or aerosol particle phases in the atmosphere, thereby advancing our understanding of the environmental fate of TPHP in diverse aqueous matrices.

### CRediT authorship contribution statement

**Yi Wu:** Writing – original draft, Investigation, Formal analysis, Data curation. **Feng Wu:** Writing – review & editing, Visualization. **Marcello Brigante:** Writing – review & editing, Supervision, Methodology, Conceptualization. **Angelica Bianco:** Writing – review & editing, Supervision, Conceptualization.

### 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 sincerely acknowledge the support from the China Scholarship Council (CSC) (Grant No. 202306270051), which is gratefully appreciated. Angelica Bianco acknowledges the French National Research Agency OPTIC (ANR-23-CE01-0015-01). The authors appreciate helpful comments from the editors and reviewers.

### Appendix A. Supplementary data

Supplementary data to this article can be found online at https://doi.org/10.1016/j.envpol.2025.126717.

### Data availability

Data will be made available on request.

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