# Re-assessment of mercury isotope mass-independent fractionation during monomethylmercury photodegradation pathways in freshwater

Elaheh Lotfi-Kalahroodi, Emmanuel Tessier, Mickael Le Béchec, Thierry

Pigot, David Amouroux

► To cite this version:

Elaheh Lotfi-Kalahroodi, Emmanuel Tessier, Mickael Le Béchec, Thierry Pigot, David Amouroux. Re-assessment of mercury isotope mass-independent fractionation during monomethylmercury photodegradation pathways in freshwater. Journal of Hazardous Materials, 2025, 494, pp.138484. 10.1016/j.jhazmat.2025.138484. hal-05396276

HAL Id: hal-05396276

https://univ-pau.hal.science/hal-05396276v1

Submitted on 3 Dec 2025

HAL is a multi-disciplinary open access
archive for the deposit and dissemination of sci-
entific research documents, whether they are pub-
lished or not. The documents may come from
teaching and research institutions in France or
abroad, or from public or private research centers.

L'archive ouverte pluridisciplinaire HAL, est
destinée au dépôt et à la diffusion de documents
scientifiques de niveau recherche, publiés ou non,
émanant des établissements d'enseignement et de
recherche français ou étrangers, des laboratoires
publics ou privés.

<figure><img src="image_1.png" /></figure>

Distributed under a Creative Commons Attribution 4.0 International License
Research Paper

# Re-assessment of mercury isotope mass-independent fractionation during monomethylmercury photodegradation pathways in freshwater

Elaheh Lotfi-Kalahroodi*, Emmanuel Tessier, Mickael Le Bechec, Thierry Pigot, David Amouroux ID

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

Université de Pau et des Pays de l'Adour, E2S/UPPA, CNRS, Institut des sciences analytiques et de physico-chimie pour l'environnement et les matériaux (IPREM), Pau 64000, France

## GRAPHICAL ABSTRACT

<figure><img src="image_3.png" /></figure>

## ARTICLE INFO

**Keywords:**

Photoreduction

Photodemethylation

Irradiation

Mercury isotopic fractionation

Gaseous mercury

Redox condition

## ABSTRACT

Photodegradation of monomethylmercury (MMHg) in surface waters induces mass-dependent and mass-independent fractionation (MIF) of mercury (Hg) isotopes, providing information on the fate of Hg. We aim to understand the mechanisms of MMHg photodegradation and Hg isotopic fractionation affected by physico-chemical factors in freshwater. We investigated the MMHg photodegradation under anoxic and oxic conditions. In addition, MMHg solution was exposed to simulated solar radiation wavelength including or not short UVB (280-305nm). Results for Hg species remaining in the solution showed lower ε¹⁹⁹Hg in anoxic conditions within the range from -8.5 ± 0.6‰ to -16.9 ± 1.9‰ and -10.8 ± 0.7‰ to -29.3 ± 6.0‰ with and without short UVB, respectively, than oxic conditions from -16.8 ± 4.0‰ to -19.6 ± 4.1‰ and -29.1 ± 4.5‰ to -33.0 ± 7.7‰ regardless of purging mode. This is probably caused by slower spin interconversion in oxic conditions. In anoxic conditions, the average Δ¹⁹⁹Hg/Δ²⁰¹Hg slope increased from 1.42 ± 0.04 to 1.67 ± 0.08 with short UVB. Our study supports the finding that Δ¹⁹⁹Hg/Δ²⁰¹Hg generated by the photochemical magnetic isotope effect (MIE) varies significantly according to the UV spectral ranges. This highlights the importance of UV light interactions for Hg and other elements featuring photochemical MIE, providing a new basis for interpreting Hg odd-MIF signature.

<footnote>* Corresponding author.</footnote>

<footnote>E-mail address: Elaheh.lotfi.kalahroodi@gmail.com (E. Lotfi-Kalahroodi).</footnote>
# 1. Introduction

Monomethylmercury (MMHg) has been recognized as a toxic pollutant because of its bioaccumulation in the food web and human exposure via seafood consumption [1–3]. To assess the MMHg exposure risks, increasing efforts are being made to understand the processes that control its speciation and, finally, biogeochemical reactions in aquatic systems. MMHg demethylation is considered a key process for removing MMHg from natural waters [4]. MMHg can be demethylated through microbial processes [5,6], and abiotic processes via photochemical reactions [7,8] and/or chemical reduction [9–12]. Since the loss of up to 80 % of MMHg by photodegradation in freshwaters [13,14] can limit MMHg uptake by aquatic organisms, MMHg photochemical process in surface waters is the major reaction. Hence, several studies have been conducted to understand the MMHg photodegradation process [15–19] and quantify it in natural waters [8,13,14,20–22]. The impact of several key factors such as light wavelength [8,13,19,21,23–25], pH [24], dissolved organic matter (DOM) type and amount [11,15,17,26–32], solar radiation with depth and DOM absorption [33–36], photoreactive trace metals [10], and radicals [12,37] on MMHg photochemical reactions have been investigated.

Understanding and determining the dominant pathways are crucial for elucidating photodegradation mechanisms to evaluate Hg retained within an ecosystem and its potential for bioaccumulation. Hg isotope fractionation is a powerful tool for quantifying sources and tracking the transformation pathways of the Hg biogeochemical cycle [38]. Although dark abiotic and biological reductions caused the enrichment of mass-dependent fractionation (MDF) δ²⁰²Hg in the remaining solution [39,40], enrichment in mass-independent fractionation (MIF) of odd isotopes; Δ¹⁹⁹Hg and Δ²⁰¹Hg was observed only following the photochemical transformation of MMHg and inorganic mercury (IHg) [16–18, 41]. This is driven by the magnetic isotope effect (MIE) [42], suggesting Hg isotope fractionation can be used as a tracer in both understanding and quantifying Hg biogeochemical pathways in aquatic ecosystems and bioaccumulation in the trophic chain. MDF is caused by differences in zero-point vibrational energies and bond strengths between isotopes which can be observed in kinetic and equilibrium processes [43,44], whereas MIF arises from the nuclear field shift (NFS) effect driven by the nuclear size and nuclear charge radius for even and odd isotopes [45, 46], and the MIE originating from unpaired nuclear spins for odd Hg isotopes [42,47].

A few laboratory experiments investigated MMHg photochemical process to demonstrate how environmental variation in DOM and solar irradiation may influence MIF of odd Hg isotopes [15–18,32]. For example, Bergquist et al., 2007 demonstrated that an increase in DOM concentration from 1 mgL⁻¹ to 10 mgL⁻¹ in the identical remaining fraction of MMHg at 0.80 leads to an increasing MIF Δ¹⁹⁹Hg extent from 0.68 ‰ to 1.69 ± ‰ [16]. In addition, Zheng et al., 2009 obtained a distinct enrichment factor (εHg¹⁹⁹) from -5.6 ‰ to -1.94 ‰ during photo-reduction with a variation of 0.002 < DOM/Hg < 0.02 because of the difference in Hg-DOM bonds influencing MIE [32]. Moreover, the presence of a reduced sulfur functional group resulted in negative MIF in the reactant [15,31], and lower εHg¹⁹⁹ were reported for a lower S<sub>red</sub>--DOM/MMHg ratio [17]. The influence of light wavelengths far lower than those present at the Earth's surface on Hg isotope fractionation during MMHg photodegradation was investigated by Malinovsky et al., 2010 who demonstrated the role of radical pairs under UVC during MMHg photodegradation [48].

Although a range for Δ¹⁹⁹Hg/Δ²⁰¹Hg slope has been previously calculated *in silico* for the NFS effect from 1.6 to 2.7 [46,49–52], no value for Δ¹⁹⁹Hg/Δ²⁰¹Hg slope has been calculated or predicted for MIE expression due to photochemical process. So far, the relation between the two MIF values of Hg odd isotopes Δ¹⁹⁹Hg and Δ²⁰¹Hg are being used to distinguish between two different photochemical processes, e.g., MMHg photodemethylation and IHg photoreduction reactions. Previous works have presented a specific and distinct relation between

Δ¹⁹⁹Hg/Δ²⁰¹Hg slopes at 1.36 ± 0.02 and 1.00 ± 0.02 for MMHg photodemethylation and IHg photoreduction, respectively [15,16,32,48, 53]. These values were introduced as references, and rapidly adopted as standards which are being used to identify the specific chemical pathways of MMHg in the environment. For example, Δ¹⁹⁹Hg/Δ²⁰¹Hg slope found in fish samples have been used to estimate the loss of MMHg via photodemethylation and IHg photoreduction to volatile gaseous mercury (Hg⁰) from surface waters to the atmosphere [16,17,54–58]. Chandan et al. has well argued the odd MIF in fish reflects a mixture of MMHg photodemethylation and IHg photoreduction leading to a lower Δ¹⁹⁹Hg/Δ²⁰¹Hg slope [17]. However, these Δ¹⁹⁹Hg/Δ²⁰¹Hg slopes were obtained by laboratory experiments performed at specific conditions, e. g., unrealistic DOM/MMHg ratio and using natural sunlight on different days, which makes difficult the judgment on variability in MIF extent and Δ¹⁹⁹Hg/Δ²⁰¹Hg slope because of seasonal variations in solar radiation. Moreover, except for Hg, significant MIE due to the photochemical mechanisms has been observed and reported for odd isotopes of several chemical elements such as ³³S in UVC and UVB photolysis of sulfone [59], ¹¹⁷Sn and ¹¹⁹Sn in photochemical reaction of methyltin under long UVC (260 nm) [60], and ⁷³Ge in UVC and UVB photolysis of methyltriphenyl germane [61]. Hence, a lack of knowledge on the specific influence of UVB light on the MIE of Hg and other chemical elements requires further investigation of odd-MIF during MMHg photochemical degradation under UVB light wavelength in aqueous media.

Our purpose in this study is, therefore, to unveil the Hg isotope fractionation caused by specific environmental parameters such as redox conditions and UVB light spectral distribution prior to the use of the Δ¹⁹⁹Hg/Δ²⁰¹Hg slope as a diagnosis to identify specific mechanisms and pathways in freshwaters. Our experimental set-up emphasizes the role of UVB light under controlled conditions, which has been barely addressed in previous studies. Indeed, we attempted to ascertain whether an increase in UVB energy and wavelength range would affect the odd MIF signature. Furthermore, the effects of redox conditions on Hg MIF during MMHg photodemethylation were investigated here for the first time to assess how variable MIF enrichment factors (εHg¹⁹⁹ and εHg²⁰¹) and the Δ¹⁹⁹Hg/Δ²⁰¹Hg slope would be affected. Hence, in our work, Hg isotopes fractionation was investigated during MMHg photodegradation in the presence of Suwannee River Fulvic Acid (SRFA) exposed to simulated solar radiation with either included short wavelength UVB (280–800nm) or excluded short wavelength UVB (305–800nm). We tried to compare laboratory experiment conditions using the continuous irradiation purging (Cont-ip) mode used in the previous works [15–18] and simulate environmental conditions using the post-irradiation purging (Post-ip) mode. Both were performed under anoxic and oxic conditions.

# 2. Material and methods

## 2.1. *Experiment set-up*

The experiment has been described in our previous work [19]. Briefly, the experiment was performed in 50 mL perfluoroalkoxy (PFA) containers by mixing 35 mL of SRFA solution at 560 µmol L⁻¹ of dissolved organic carbon (DOC) with MMHg solution at 0.25 µmol L⁻¹. The initial pH of the solution was adjusted to pH 6.5–7 using ultrapure HCl and NaOH. Reactors were wrapped in aluminum foil and stirred overnight in a dark room to equilibrate MMHg complexation with DOM. For the first purging mode, the continuous irradiation purging (Cont-ip) experiment, solutions were continuously purged during the irradiation using argon (Ar) and/ or pure air a with the adequate gas flow (Fig. S1), calculated according to assuming Henry's law equilibrium between Hg concentration in the equilibrated headspace and the solution using irradiation time to obtain > 99 % purging efficiency of Hg⁰ based on the irradiation time [62]. For the second purging mode, post-irradiation purging (Post-ip) experiment (Fig. S1): i) reactors were then purged
with Ar, and/or pure air (previously cleaned by set-up a gold trap) for 10 min with a gas flow at 100 mL min⁻¹ to produce anoxic and oxic conditions, respectively, ii) solutions were then irradiated for 3, 10, 30, and 60 min, and iii) solutions were again purged by Ar for 10 min at 100 mL min⁻¹ to collect Hg⁰ on a trap. The Hg⁰ was collected through traps made of a quartz glass tube filled with gold-coated silica (Amasil™ with a trapping capacity of 82 mg of Hg, a recovery > 97% at 30 °C <T < 210 °C), which were purchased from P S Analytical. The 10 mL of remained solutions were acidified at 1 % HCl (v/v) for Hg speciation, and the rest was digested by the addition of 1 % monobrominchloride (BrCl) (v/v) for measurement of Hg isotopic composition. To neutralize the excess of BrCl, 0.5% (v/v) of a 30% solution of hydroxylamine hydrochloride (NH₂OH.HCl) was added to the solution before the introduction of the sample to the instrument [63]. All the samples were stored in the dark at 4 °C before analysis. The solutions were irradiated using a polychromatic light source (Oriel Instruments with a 300 W Xe arc lamp) with a beam-turning Mirror and full reflector (Newport 66215). Light emission to solutions was directed through a cutout window on an aluminum sheet to avoid the uncontrollable emission of light from surfaces. The distance between the solutions and the light source was 20.5 cm. The experiment was performed under two wavelength ranges “+short UVB” (280–800 nm, irradiance (I<sub>simulator</sub> at 29 mW cm⁻²)), and “no short UVB” (305–800 nm, I<sub>simulator</sub>= 27 mW cm⁻²) using a 65CGA-305 filter (GB328 Newport Corporate Irvine California USA). Irradiances of simulator and natural sunlight were measured using an AVASPEC 2048XL spectroradiometer equipped with a 600µm diameter optical fiber and a 4.9 mm diameter cosine corrector (Fig. S2). These data illustrate the choice to use the high intensity of simulator irradiance under a short UVB wavelength range (280–305 nm) compared to that for natural sunlight (I<sub>sunlight</sub>), I<sub>simulator</sub>: I<sub>sunlight</sub> at 38:1 [19]. All the experiments were performed in triplicate, and experimental uncertainty was calculated by triplicate samples’ standard errors (2SE) [22,63–66]. Three series of control experiments were performed in the dark (wrapped reactors using aluminum foil), without DOM, and exposure to visible light (400–800 nm, I<sub>simulator</sub>= 12 mW cm⁻²) using a 65CGA-400 filter (CA222) in the same conditions as described above (Table S1). The irradiation times (s) were converted to absorbed energy (J) using measured simulator irradiance (W cm⁻²), absorbance of initial SRFA solution, and the reactor’s physical parameters based on a previous work [67] (Supplementary Information).

## 2.2. Chemical products

All acids used in this work were trace metal or optima-grade quality. The 50 mL PFA containers used were cleaned with 10% (v/v) HNO₃ twice, and 10% (v/v) HCl, then deionized H₂O. All these steps were carried out overnight and for 1 h in the ultrasonic bath, and finally dried under a laminar flow hood overnight. All solutions were prepared using ultrapure water 18 MΩ (Milli-Q system, Millipore). Analytical grade NaOH and ultrapure HCl (J.T. Backer) were used for the sample preparation. A stock solution of MMHg (50 µmol L⁻¹ as Hg) was prepared from methylmercury chloride (Strem Chemicals, Inc. Bischheim, France) in ultrapure acetic acid/ methanol with a ratio of 3:1 (v/v). The DOM stock solution (650 µmol L⁻¹ as DOC) was made from SRFA II (ref. 2S101F) purchased from the International Humic Substances Society (IHSS with C = 52.3 %, H = 4.4 %, O = 43 %, S = 0.4 %, and N = 0.7 % as a mass fraction) in ultrapure water. The chemical element concentrations and low molecular weight molecules detected in SRFA have been described in Tables S2 and S3 [68–71]. The pH of this solution was adjusted to 6.5–7, and it was filtered through a 0.2 µm pore-size nylon membrane (VWR®) using a syringe filter to prevent the participation of particulate organic matter in Reactive Oxygen Species (ROS) generation [72].

## 2.3. Quantification of mercury species

The quantification of MMHg and IHg in remaining solutions was performed by double species-specific isotopic dilution method (D-SS-IDA) and analyzed by gas chromatography- inductively coupled plasma mass spectrometry (GC-ICPMS) [73,74]. Briefly, before analyses, samples were derivatized at pH 3.9 by ethylation using sodium tetraethylborate (NaBEt₄, 5%), then extracted in isooctane by shaking using an orbital shaker for 20 min. Trace Ultra GC equipped with a Triplus RSH autosampler coupled to an ICP-MS XSeries II (Thermo Scientific, USA) [75–77]. Hg⁰ collected on Amasil traps was analyzed through an atomic fluorescence spectrometry (AFS) using a Hg analyzer (PSA 10.525 Sir Galahad II), as described in previous work [78]. The mass balance was calculated through the sum of the Hg species (MMHg, IHg, and Hg[¹⁰¹H]) and compared to the initial total MMHg. Mass balance calculation revealed an error of < 9% for the initial total MMHg.

## 2.4. Determination of mercury isotope fractionation

The analysis was performed with a multicollector-inductively coupled plasma-mass spectrometer (MC-ICP-MS, Nu Instrument) using introduction with cold vapor generation (CVG) [63]. A desolvation nebulization (DSN-100, Nu Instrument) system helped to introduce a standard reference material of thallium (NIST-997) simultaneously to correct instrumental mass bias through the exponential fractionation law [64]. A standard-sample bracketing system was used to calculate Hg isotopic composition (δ⁵Hg(‰), x = 199–204) relative to the standard reference material of IHg (NIST-3133, n = 291). The isotopic composition δ⁵Hg is used to expressed:

$$ \delta^x \text{Hg}(\text{\textperthousand}) = 10^3 \left[ \frac{(^{x}\text{Hg}/^{198}\text{Hg})_{\text{unkown}}}{(^{x}\text{Hg}/^{198}\text{Hg})_{\text{NIST-3133}}} - 1 \right] \quad (1) $$

Where xHgA⁹⁸HgNIST-3133 is the average isotopic ratio of two bracketing standards. Deviation from the theoretical mass-dependent fractionation (MDF, δ²⁰²Hg) line presents the mass-independent fractionation (MIF) which is expressed using capital delta (ΔxHg) [16, 79]:

$$ \Delta^x \text{Hg} = \delta^x \text{Hg} - (\delta^{202} \text{Hg} \times \beta_{\text{kin}}) \quad (2) $$

$$ \beta_{\text{kin}} = \ln(m_{198}/m_{x})/\ln(m_{198}/m_{202}) \quad (3) $$

The external reproducibility is evaluated using a secondary reference material, NIST RM 8610 (Mercury Isotopes in UM-Almaden Mono-Elemental Secondary Standard), which was analyzed in each analytical session at 5 nmolL⁻¹ and measured at -0.53 ± 0.10 ‰ for δ²⁰²Hg, -0.03 ± 0.09 ‰ for Δ¹⁹⁹Hg and -0.04 ± 0.09 ‰ (2 SD, n = 51) [66] for Δ¹⁹⁹Hg, in line with the previous studies in our laboratory [22,63–65].

## 2.5. Kinetic fractionation factor

By assuming kinetic isotope fractionation, the photodegradation process can follow a Rayleigh distillation model (Fig. S2). The fractionation factors can be calculated based on δ⁵Hg results of the reactant i.e. MMHg as a larger pool compared to IHg during photodemethylation using the Rayleigh distillation equation defined by Mariotti et al. [80]:

$$ (\alpha_{\text{Hg}0}^{x} / \text{MMHG} - 1) \ln f = \ln \frac{10^{-3} \delta_{\text{Hg}0}^{x} + 1}{10^{-3} \delta_{\text{Hgint}}^{x} + 1} \quad (4) $$

Where xαHg⁰/MMHg is the fractionation factor for the photodemethylation reaction, f is the fraction of MMHg and IHg remaining in the solution at time t, δ⁵Hginitial and δ⁵Hgt are the isotopic composition of the initial solution (before irradiation) and the remaining solution at time t, respectively. The enrichment factor ε⁵Hg⁰/MMHg (‰) of photodemethylation was then calculated using:
$$ \epsilon_{Hg^0/\text{MMHg}}^x = 10^3 \left( \alpha_{Hg^0/\text{MMHg}}^x - 1 \right) \qquad (5) $$

By rewriting two equations above:

$$ \left(10^{-3} \epsilon_{Hg^0/\text{MMHg}}^x\right) \ln f = \ln \frac{10^{-3} \delta_{Hg^0}^x + 1}{10^{-3} \delta_{Hg\text{int}}^x + 1} \qquad (6) $$

the slope of the best linear fit of $10^3 \times \ln [(10^{-3} \times \delta^x \text{Hgt} + 1) / (10^{-3} \times \delta^x \text{Hginitial} + 1)]$ as a function of $\ln f$ (MMHg+IHg) (Eq. (6)) according to York regression [81] using OriginLab 2020 as previous work [22], presents the enrichment factor $\epsilon^x \text{Hg}^0/\text{MMHg}$, which can be used to calculate kinetic fractionation factors $\alpha^{202}$ and $\alpha^{199}$ (Eq. (5), Table S4) [16,18,22,48]. Similarly, $\Delta^{199} \text{Hg}/\Delta^{201} \text{Hg}$ slope were determined from the best linear fit of $\Delta^{199} \text{Hg}$ versus $\Delta^{201} \text{Hg}$ by York regression [81] using OriginLab 2020 as previous work [22], which is more accurate than calculating an average ratio when MIF is close to zero. The errors associated with the slope are 2SE, as reported in Table S5.

## 2.6. Mercury chemical desorption and sparging

To measure the isotopic signature of Hg[0], in the Cont-ip experiment in anoxic and oxic conditions, Hg⁰ collected on Amasil traps was sparged in an oxidizing solution according to a previous method [82]. Sparging was performed for duplicate samples irradiated for 10 and 60 min. A thermal desorption system was used to desorb Hg⁰ from the Amasil trap, and then lead it into an oxidizing solution. The system consists of two nickel-chrome wire heaters wrapped around a quartz tube connected with a programmable DC power supply to control the temperature in the quartz tube and the desorption process (Fig. S3). Hg⁰ loaded on Amasil traps was slowly heated step-wise from room temperature to 525°C over 33 min in 4 temperature intervals. Hg⁰ released from the Amasil trap was conducted into 6 mL oxidizing solution reverse-aqua regia at 40 % HNO₃:HCl (3:1) through an Ar flow at 3–4 mL min⁻¹. Sparging solutions were analyzed for total Hg using GC-ICP-MS as described above and then stored at 4°C for Hg isotopic composition measurement. The total Hg concentration in the blanks never exceeded 0.32 nmolL⁻¹, and recovery was determined at 104 ± 4 %. The mass balance and error propagation calculated for Hg⁰ reflected a complementary reservoir of the Hg species (MMHg+IHg) remainin in the solution (Supplementary Information file Eq. 3 and 4).

# 3. Results & discussion

## 3.1. Redox conditions effects on mercury isotopic compositions

We studied the simultaneous MMHg photodemethylation to IHg and IHg photoreduction to Hg⁰ under both anoxic and oxic conditions by irradiation of MMHg in a model freshwater containing dissolved humic substances. The solutions were exposed to simulated solar radiation including short wavelength UVB (280–305 nm), hereafter called "+short UVB" (280–800 nm), or excluding short wavelength UVB (> 305 nm), hereafter called "no short UVB" (305–800 nm), since UVB distribution is not identical from the northern hemisphere to tropical regions. In addition to Hg speciation, we determined the isotopic composition of the remaining Hg species, e.g., MMHg and IHg in the solution for all experiments and Hg⁰ for the Cont-ip experiment. We then calculated the enrichment ($\epsilon_{Hg}^x$), fractionation ($x \alpha_{Hg}^0/\text{MMHg}$) factors, and MIF slopes ($\Delta^{199} \text{Hg}/\Delta^{201} \text{Hg}$). Since no evidence of methylation and demethylation were observed in the dark control experiment [19], no significant variation in MDF ($\delta^{202} \text{Hg}$) and MIF ($\Delta^{199} \text{Hg}$ and $\Delta^{201} \text{Hg}$) was measured (Table S1). All isotopic data are presented in the Supplementary Information file. Moreover, in the additional control experiments without DOM and under visible light (400–800 nm) insignificant variation in both MDF and MIF was observed (Table S1). Insignificant variation in MMHg photodegradation has been already reported in visible light experiments for both redox conditions and purging modes in our previous

work [19]. In all experiments, MMHg photodemethylation and IHg photoreduction to Hg⁰ resulted in variation in both MDF and MIF (Fig. 1, & Tables S6 & S7). Our previous work on the kinetic of the MMHg photochemical process has demonstrated that MMHg photodecomposition to Hg⁰ is mainly induced by a first photodemethylation step to IHg followed by an irreversible photoreduction step to Hg⁰ and Hg⁰ oxidation is negligible [19]. Overall, all the MMHg photodegradation experiments displayed kinetic MDF of Hg isotopes with enrichment of heavier isotopes in the reactant, meaning the lighter isotopes were preferentially reduced to Hg⁰ (Fig. 1, Tables S6 & S7), consistent with kinetic isotope fractionation and previous work [16–18].

The MDF increased from an initial value at −0.73 ± 0.10 (2SE) ‰ to a final value ranging from −0.54 ± 0.10 ‰ to −0.13 ± 0.15 ‰ (Tables S6 & S7). For the whole of the experiments, an overall positive trend is observed showing the reactant enriched in $^{202} \text{Hg}$ for 0.60 ± 0.18 (2SE)‰, 0.36 ± 0.12 ‰ in anoxic, and 0.45 ± 0.12 ‰, and 0.27 ± 0.10 ‰ in oxic conditions under + short UVB and no short UVB in the Cont-ip experiment, respectively compared to the $^{202} \text{Hg}$ value in the initial solution (Fig. 1 and Table S6). Similar results were obtained in the Post-ip experiment (Fig. 1 and Table S7). Overall, the odd MIF was more influenced by the MMHg photochemical process than MDF. Under all conditions, the odd MIF followed Rayleigh-type fractionation, meaning that with progress in the photochemical process, the remaining solution was enriched by heavier odd Hg isotopes (Fig. S4 & Tables S6 & S7). The direction of odd MIF displayed positive MIE (described in Section 3.3) as the underlying cause for odd isotopes enrichment in reactants, up to 3.15 ± 0.07 (2SE, $f = 0.84$) ‰ for $\Delta^{199} \text{Hg}$, and 2.06 ± 0.15 ‰ for $\Delta^{201} \text{Hg}$ in anoxic + short UVB in the Cont-ip experiment. Blocking short UVB wavelength reduced MIF signatures $\Delta^{199} \text{Hg}$ to 2.14 ± 0.44 ‰, and $\Delta^{201} \text{Hg}$ to 1.52 ± 0.32 ‰, and removing all UV wavelengths suppressed the photodemethylation [19] and MIF signature variation (Table S1). This relationship between the MIF magnitude and the light wavelength has been demonstrated in previous works [16,18]. Rose et al., 2015 demonstrated that removing UVB (290–320 nm) suppressed MIF signatures from 0.55 ± 0.10 ‰ to 0.17 ± 0.00 ‰ [18]. They observed the most pronounced MIF signature in the remaining solutions exposed to UVB and, therefore, concluded that despite the low effect of light intensity on MMHg loss and photodemethylation, light wavelength distribution has a high impact on the odd MIF signatures [18]. The measurement of the Hg isotopic composition of produced Hg⁰ from MMHg photodegradation in the Cont-ip experiment demonstrated the enrichment of Hg⁰ in lighter Hg isotopes (Tables S8 & S9). The mass balance, error propagation calculation (Table S8 & S9, Eq. 3 & 4 in Supplementary Information file) and comparison with the initial Hg isotopic composition validated Hg⁰ as a complementary reservoir of Hg species remaining in the solution. As assumed in previous work [16], our data demonstrated a negative MIF $\Delta^{199} \text{Hg}$ for the complementary reservoir Hg⁰ mirroring the positive MIF of Hg species in the solution during MMHg photodegradation. Our results are consistent with previous work where Bouchet et al. reported the negative odd MIF for IHg down to −9.9 ± 0.2 ‰ as the product of MMHg photodemethylation in mesotrophic and eutrophic lake waters incubated under various redox conditions [22].

In our work, we highlighted the relationship between the MIF extent and redox conditions for the MMHg photochemical process, suggesting MIE as the dominant mechanism. Our results showed that $\Delta^{199} \text{Hg}$ presents a higher MIF value in anoxic at 3.15 ± 0.07 (2SE, $f = 0.84$) ‰ with lower $f$ extent than that in oxic conditions at 1.69 ± 0.38 (2SE, $f = 0.92$) ‰ under + short UVB, and at 2.14 ± 0.644 ($f = 0.93$) ‰ than that in oxic conditions at 1.61 ± 0.17 ($f = 0.95$) ‰ under no short UVB in the Cont-ip experiment (Fig. 1 and Table S6). Similar results were again observed in the Post-ip experiment (Fig. 1 and Table S7). As expected from our discussion above, these variations in MIF signatures between redox conditions were more pronounced under + short UVB than under no short UVB in both purging modes. For both purging modes, the largest MIF was observed in the experiments exposed to anoxic
<figure><img src="image_4.png" /><figcaption>**Fig. 1.** The yield of Hg species (MMHg & IHg) remaining in the solution vs. absorbed energy, (A & D, the mass-dependent fractionation δ<sup>202</sup>Hg (B & E), and mass-independent fractionation Δ<sup>199</sup>Hg (C & F) vs. the fraction of Hg species (MMHg & IHg) remaining in the solution during MMHg photodegradation under + short UVB, and no short UVB in the continuous irradiation purging (Cont-ip) and the post-irradiation purging (Post-ip) experiments.</figcaption></figure>

conditions, whereas solutions photo-irradiated in oxic conditions
showed almost lower MIF. However, for a given *f* extent, higher MIF
signatures were obtained in oxic conditions than anoxic conditions,
regardless of light wavelength and purging mode (Fig. 1C and F). This
fact raises the question of the perceptible mechanism of Hg isotopes
under different redox conditions. Since MIE depends on the influence of
reaction rates on the nuclear magnetic moment of reactants [83], it is
worth noting that the rate-limiting step results in MIF. The rate-limiting
step involves an excited radical pair intermediate where hyper-fine
interaction (HFI) influences the fate of the radical pair intermediate.
For instance, important factors such as the lifetime and coupling of the
radical pair, and the solvent cage can affect MIE expression and thus
induce MIF [42,84–86]. Oxic conditions exhibit lower MMHg photo-
demethylation and photoreduction rate constants, measured respec-
tively at 4.5 ± 0.4 kJ<sup>-1</sup> and 2.8 ± 0.2 kJ<sup>-1</sup>, compared to anoxic
conditions determined at 18.0 ± 2.2 kJ<sup>-1</sup> and 11.1 ± 0.5 kJ<sup>-1</sup>, respec-
tively [19]. This supports higher activation energy under oxic conditions
and faster intersystem crossing (ISC) of odd isotopes affected by MIE
than in anoxic conditions. The MIE is observed when different magnetic
and non-magnetic nuclei are distributed between the reaction products
[42,78]. For example, in oxic conditions, competitive reactions caused
by O<sub>2</sub> or the production of a less stable triplet state may generate faster
spin interconversion between singlet and triplet states [87], which leads
to higher enrichment of odd isotopes in oxic conditions than in anoxic

conditions in the reactant (i.e., MMHg). Indeed, for a given MMHg
photodegraded fraction, a greater MIF extent is expected in an oxic
condition than in an anoxic condition for the remaining solution, despite
a higher photodemethylation rate constant in anoxic conditions [19].
This supports that the oxic conditions contribute more to MIF signatures
compared to anoxic conditions and affects strongly odd isotopes
although the reactions progress slowly.

3.2. UVB distribution and redox condition influences on odd MIF

Despite the larger variation in the extent of MDF in the Cont-ip
compared to the Post-ip experiment (Fig. 1 and Tables S6 and S7),
δ<sup>202</sup>Hg fractionation factors (²⁰²α) stretched over an order of magnitude
relatively constant from 0.9978 ± 0.0007–0.9895 ± 0.0089 for all the
experiments (Table S4). As discussed above and as a result of lower MIF
extent, anoxic photodemethylation displayed a lower isotopic fraction-
ation factor at ¹⁹⁹α = 0.9915 ± 0.0008 than oxic photodemethylation at
¹⁹⁹α = 0.9832 ± 0.0040 under + short UVB in the Post-ip experiment
(Table S4). Our results showed a similar trend under no short UVB
experiment, meaning a lower isotopic fractionation factor in an anoxic
photochemical process ¹⁹⁹α = 0.9892 ± 0.0011 obtained for a higher
photodemethylation and photoreduction rate constants in anoxic con-
ditions (8.6 ± 0.5 kJ<sup>-1</sup> and 4.9 ± 0.2 kJ<sup>-1</sup>) [19]. Unlike the Post-ip
experiment, ¹⁹⁹α demonstrated similar data in the Cont-ip experiment
with no discernible trend between diverse parameters (Table S4). Odd isotopes were highly impacted and fractionated under oxic conditions compared to anoxic conditions in the Post-ip experiment, while removing photoreactive volatile organic carbon (VOC) from the solution [19] in the Cont-ip experiment showed an insignificant effect on the isotopic fractionation factor. A previous work obtained a lower average value of isotopic fractionation factor at 199α = 0.9940 for MMHg photodegradation [16] than that in our results (199α= 0.9723, Table S4) in similar conditions, meaning under + short UVB in oxic conditions of the Cont-ip experiment. This last high value may be related to photodemethylation rate constants, as 20 % of MMHg was photodegraded after 300- and 30-min irradiation in Bergquist et al., 2007 work [16] and our experiment, respectively. Indeed, photodemethylation and photoreduction rate constants are controlled by experimental parameters, including DOM type, concentration [11,15,17,26–32], and intensities of light irradiance [8,13,16,18,19,21,23–25].

The data followed Rayleigh fractionation, allowing the calculation of the enrichment factors of ε<sub>Bg</sub><sup>202</sup>, ε<sub>Bg</sub><sup>199</sup>, and ε<sub>Bg</sub><sup>201</sup> (Fig. S6 and Table S10). As expected, MDF enrichment factors (ε<sub>Bg</sub><sup>202</sup>) showed lower values in anoxic conditions at -2.2 ± 0.7 ‰ (2SE, York regression, R<sub>Pearson</sub> = 0.88), and -8.2 ± 9.1 (R<sub>Pearson</sub> = 0.46) than those within the same range at -7.1 ± 5.6 ‰ (R<sub>Pearson</sub> = 0.47) and -6.0 ± 1.4 ‰ (R<sub>Pearson</sub> = 0.92) in oxic conditions in + short UVB and no short UVB in the Post-ip experiment, respectively (Table S10). Similarly, lower ε<sub>Bg</sub><sup>199</sup> was observed in anoxic conditions at -8.5 ± 0.6 ‰ (R<sub>Pearson</sub> = 0.99), and -10.8 ± 0.7 ‰ (R<sub>Pearson</sub>=1.00) than those at -16.8 ± 4.0 ‰ (R<sub>Pearson</sub> = 0.93) and -33.0 ± 7.7 ‰ (R<sub>Pearson</sub> = 0.99) in oxic conditions in + short UVB and no short UVB in the Post-ip experiment, respectively. ε<sub>Bg</sub><sup>199</sup> obtained under + short UVB in oxic condition in the Cont-ip at -19.6 ± 4.1 ‰ are consistent with average ε<sub>Bg</sub><sup>199</sup> values at -14.9 ± 1.2 ‰ reported during MMHg photodemethylation in the presence of various MMHg: S<sub>red-DOM</sub> as performed in the similar condition [17]. In addition, ε<sub>Bg</sub><sup>199</sup> value reported at -14.3 ± 0.6 ‰ for the in situ lake waters incubation experiment by Bouchet et al. [22] is within the ranges of ε<sub>Bg</sub><sup>199</sup> at -16.8 ± 4.0 ‰ and -16.9 ± 1.9 ‰ calculated for + short UVB with the oxic condition in the Post-ip experiment and anoxic conditions in the Cont-ip experiment, respectively. These comparisons showed that those fractionation extents are consistent with the fact that similar photodegradation pathways occur in the field and laboratory experiments.

In previous work, we measured the role of ROS in the photo-
demethylation pathways using three scavengers, including isopropanol,
isoprene, and sodium azide as probes of hydroxyl radicals, triplet excited
state dissolved natural organic matter, and singlet oxygen, respectively,
in oxic and anoxic conditions for the Post-ip experiment [19]. The
decrease of specific UV absorbance in solutions in oxic compared to
anoxic conditions confirmed DOM photobleaching in oxic conditions
[19]. That study proved that oxygen inhibits and modifies the photo-
demethylation and photoreduction kinetic pathways [19]. At the same
time, it enhances the production of ROS [9–12,88] and DOM photo-
bleaching [19,89,90]. In the Post-ip experiment, in oxic conditions, the
distinct difference in MIF enrichment factors at ε<sub>Bg</sub><sup>199</sup> = -16.8 ± 4.0 ‰,
and -33.0 ± 7.7 ‰ under +short UVB and no short UVB, respectively,
compared to those values at -8.5 ± 0.6 ‰, and -10.8.0 ± 0.7 ‰ in
anoxic conditions (Fig. 2 and Table S10), highlights that the MMHg
photodegradation mechanism in oxic conditions is highly affected by
UVB spectral distribution. In contrast, it is more independent of light
wavelength in anoxic conditions. By removing VOC in the Cont-ip
experiment, this trend was not observed between redox conditions.
However, higher ε<sub>Bg</sub><sup>199</sup> values in + short UVB at -16.9 ± 1.9 ‰ and -19.6
± 4.1 ‰ than those in no short UVB -29.3 ± 6.0 ‰ and -29.1 ± 4.5 ‰ in
anoxic and oxic conditions, respectively, demonstrated that the UVB
spectral distribution differentiated between enrichment factors ε<sub>Bg</sub><sup>199</sup>
(Fig. 2). Notably, the prominence of UVB spectral distribution compared
to redox conditions in the Cont-ip experiment has been previously
proven by reporting the MMHg photodemethylation rate constants in
+ short UVB within the same range at 1.2 ± 0.5 kJ<sup>-1</sup> and 1.8 ± 1.0 kJ<sup>-1</sup>

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

<figure><img src="image_6.png" /><figcaption>Fig. 2. Ln(Δ<sup>199</sup>Hg) calculated according to Eq. (6) vs. the fraction of Hg species (MMHg & IHg) remaining in the solution during MMHg photodegradation under + short UVB and no short UVB in the continuous irradiation purging (Cont-ip, A) and the post-irradiation purging (Post-ip, B) experiments.</figcaption></figure>

in anoxic and oxic conditions [19]. However, the exponential curve
fitting displayed a significant relationship (R² = 0.76 and 0.80, Fig. 3)
between ε<sub>Bg</sub><sup>199</sup> and photodemethylation and photoreduction rate con-
stants, respectively. These relations support that odd-MIF fractionation
factor 199α is clearly affected by reaction rate constants, which are in
turn controlled by short UVB spectral distribution and aqueous redox
conditions. As discussed above, high activation energy and fast inter-
system crossing of odd isotopes affected by MIE occurred for the reaction
with lower rate constants. As a consequence, a high isotopic fraction-
ation factor and high odd-MIF of Hg isotopes are observed for low
photodemethylation and photoreduction rate constants (e.g. in oxic
conditions). Our results also confirmed the influence of redox conditions
on the odd-MIF of Hg isotopes, probably reflecting distinct mechanisms
and reaction pathways.

3.3. Impact of UVB spectral distribution on MMHg photodegradation isotopic proxy

The average Δ¹⁹⁹Hg/Δ²⁰¹Hg slope of the residual MMHg and IHg in the solution exposed to + short UVB calculated at 1.69 ± 0.06 (2SE, York linear regression 95 % confidence [81], R<sub>Pearson</sub>= 0.99) in anoxic and 1.68 ± 0.18 (R<sub>Pearson</sub>= 0.95) in oxic conditions in the Post-ip experiment (Fig. 4B and Table S5), and at 1.64 ± 0.03 (R<sub>Pearson</sub>= 1.00), and 1.73 ± 0.09 (R<sub>Pearson</sub>= 0.98) in the Cont-ip experiments, respectively (Fig. 4A and Table S5). Although these latter slopes are close to the typical Δ¹⁹⁹Hg/Δ²⁰¹Hg values determined for NFS effect at 1.6–2.7 [46,49–52], insignificant variation in Δ²⁰⁰Hg and Δ²⁰⁴Hg confirmed that the dominant MIF of odd isotope was governed by the MIE (Tables S6 and S7). This is consistent with previous works where NFS was not the dominant isotope effect in the MMHg photochemical
<figure><img src="image_7.png" /><figcaption>**Fig. 3.** Enrichment factor (ε<sup>199</sup>Hg vs. MMHg photodemethylation (K<sub>PD</sub>, black solid line) and IHg photoreduction (K<sub>PR</sub>, grey dashed line) rate constants [19] with respect to absorbed energy under + short UVB, and no short UVB in the continuous irradiation purging (Cont-ip) and the post-irradiation purging (Post-ip) experiments.</figcaption></figure>

process [15–18,32,91].

The Δ¹⁹⁹Hg/Δ²⁰¹Hg slope dropped to lower values at 1.50 ± 0.04 (R<sub>Pearson</sub>= 1.00) and 1.49 ± 0.05 (R<sub>Pearson</sub>= 1.00) in anoxic and oxic conditions when solution irradiated under no short UVB wavelength in the Post-ip experiment, and at 1.45 ± 0.06 (R<sub>Pearson</sub>= 0.99) and 1.36 ± 0.04 (R<sub>Pearson</sub>= 0.99) in the Cont-ip experiment, respectively (Fig. 4, and Table S5). MIF data of all Hg species MMHg and IHg as reactant and Hg⁰ as product displayed higher average Δ¹⁹⁹Hg/Δ²⁰¹Hg slope in + short UVB at 1.51 ± 0.02 (York linear regression 95 % confidence [81], R<sub>Pearson</sub> = 1.00) than in no short UVB at 1.41 ± 0.01 (R<sub>Pearson</sub> = 1.00) (Fig. S5), confirming noticeable MIF expressions resulted from distinct MIE mechanisms depending on short UVB spectral distribution.

Different Δ¹⁹⁹Hg/Δ²⁰¹Hg slopes were reported for the MMHg photodemethylation process depending on various conditions applied during experiments. For example, Bouchet et al., 2022 obtained a lower Δ¹⁹⁹Hg/Δ²⁰¹Hg slope at 1.24 ± 0.03 for in situ MMHg photodemethylation incubation of a solution at 3 mgL⁻¹ DOC, and in fish samples collected in the same water as characteristic of MMHg found in the surrounding environment [22], distinct from the Δ¹⁹⁹Hg/Δ²⁰¹Hg at 1.31 ± 0.01 generally considered for fish samples [16,17,38]. Besides, Chandan et al. showed the Δ¹⁹⁹Hg/Δ²⁰¹Hg slope during MMHg photodegradation varies with DOM type and concentration from 1.38 ± 0.02–1.17 ± 0.04 for low and high MMHg: S<sub>red</sub>-DOM experiments [17], although Bergquist et al., 2007 reported a Δ¹⁹⁹Hg/Δ²⁰¹Hg slope at 1.36 ± 0.02 [16]. Moreover, Rose et al., 2015 reported the average Δ¹⁹⁹Hg/Δ²⁰¹Hg slope for MMHg photodegradation at 1.33 ± 0.14 [18]. These authors concluded Δ¹⁹⁹Hg/Δ²⁰¹Hg slope remained consistent regardless of the incident radiation energy during exposure i.e. in September or July. However, a calculation of the Δ¹⁹⁹Hg/Δ²⁰¹Hg slope from their data file by York regression showed a significantly distinct Δ¹⁹⁹Hg/Δ²⁰¹Hg slope, higher in September at 1.40 ± 0.09 than that in July at 1.22 ± 0.07 [18]. In addition to experimental work, theoretical approaches showed distinct magnetic isotope fractionation, depending

on the HgX₂ and MMHgX complex identity [92]. These authors reported no MIE during the photolysis of HgX₂ and MMHgX bound to halides when Hg ligands are heavier elements, e.g., bromide and iodide. Their results are, therefore, in agreement with our conclusion indicating that the Δ¹⁹⁹Hg/Δ²⁰¹Hg slope diverges from 1.36, adopted as a reference value for MMHg photodegradation, depending on the energy of incident radiation in various environmental conditions. However, our work showed that the experimental average of Δ¹⁹⁹Hg/Δ²⁰¹Hg slopes increased from 1.42 ± 0.04 (R<sub>Pearson</sub>= 1.00) to 1.67 ± 0.08 (R<sub>Pearson</sub>= 0.99) by the expansion of the UVB radiation light spectra to shorter wavelength (Fig. S6), confirming that UVB is highly relevant to the mechanism of MIF expression in the environment.

### 3.4. Implication

This study aims to unveil the Hg isotope effects caused by specific environmental parameters such as redox conditions and UVB spectral distribution. We tried to compare laboratory experiment conditions used in the previous works, e.g., continuous irradiation purging (Cont-ip) mode, and simulate environmental conditions, e.g., post-irradiation purging (Post-ip) mode. A simple comparison between redox conditions in the Post-ip experiment results, under the same wavelength, i.e. + short UVB with ¹⁹⁹α variation from 0.9915 in anoxic to 0.9832 in oxic conditions (Table S4), revealed that various environmental parameters such as redox conditions and UV spectral distribution induce dissimilar fractionation factors. This leads to different mechanisms of MMHg photodegradation and distinct estimations of the fate of Hg within the same ecosystem. Since the estimation given in the literature varies between sites, field and experimental works require the consideration of environmental parameters such as short UVB light spectral distribution, especially for measurement and reporting Δ¹⁹⁹Hg/Δ²⁰¹Hg slopes, and redox condition in waters as limiting factor for MMHg photochemical process progress inducing different MIF signatures in natural waters. We
<figure><img src="image_8.png" /><figcaption>Fig. 4. Mass-independent fractionations of individual Δ¹⁹⁹Hg vs. Δ²⁰¹Hg during MMHg photodegradation under + short UVB, and no short UVB in the continuous-irradiation purging (Cont-ip, A) and the post-irradiation purging (Post-ip, B) experiments.</figcaption></figure>

should therefore consider Hg isotope composition measured in natural samples as an effective variable impacted by all competitive chemical processes, and environmental parameters. However, our work supports Hg MIF signatures as promising tools for helping to quantify MMHg photochemical processes. Still, this work emphasizes the need to conduct lab experiments on MMHg binding environment e.g. thiol by reinforcing MMHg photodegradation [11,29,93,94] and salinity (chloride) or conductivity (ionic strength) to find the link between MMHg photochemical reactions and MIF expression to evaluate this tool before fully utilized in natural waters. Moreover, our results reveal a wider implication of photochemical MIF of odd isotopes and highlight new research lines that include several chemical elements featuring MIE during photochemical processes.

## Environmental Implication

We aimed to unveil the mercury isotope effects caused by specific environmental parameters like redox conditions and UVB spectral distribution. A comparison between redox conditions under the same wavelength revealed that various environmental parameters can induce dissimilar fractionation factors. This leads to different mechanisms of monomethylmercury photodegradation and distinct estimations of the

fate of Hg within the same ecosystem. This comparison supports that the estimation given in the literature and our work varies between sites. They require the consideration of environmental parameters like short UVB light spectral distribution, especially for measurement and reporting ratios Δ¹⁹⁹Hg/Δ²⁰¹Hg in field data, and redox water oxygenation factor as limiting factor for photochemical process progress.

## CRediT authorship contribution statement

**Mickael Le Bechec:** Writing – review & editing, Resources, Methodology, Investigation, Formal analysis, Data curation.
**Emmanuel Tessier:** Writing – review & editing, Validation, Resources, Methodology, Investigation, Formal analysis, Data curation, Conceptualization.
**David Amouroux:** Writing – review & editing, Supervision, Resources, Project administration, Investigation, Funding acquisition, Conceptualization.
**Thierry Pigot:** Writing – review & editing, Resources, Methodology, Investigation.
**Elaheh Lotfi Kalahroodi:** Writing – review & editing, Writing – original draft, Validation, Resources, Methodology, Investigation, Formal analysis, Data curation, Conceptualization.
Declaration of Competing Interest

The authors declare the following financial interests/personal re-
lationships which may be considered as potential competing interests:
David Amouroux reports financial support was provided by Agence
National de la Recherche (ANR) (project-ANR-17-C34-0014). David
Amouroux reports financial support was provided by Swiss National
Science Foundation (SNF) (Project N 175721). If there are other authors,
they 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

We thank the Agence National de la Recherche (ANR) (project-ANR-17-C34-0014) and the Swiss National Science Foundation (SNF) (Project N 175721) for financial support in the framework of the PHYTAMBA project (PRCI ANR/SNSF) and financial support from E2S /UPPA (Energy and Environment Solutions) in the framework of the MESMIC HUB (Metals in Environmental System Microbiology, Energy and Environment Solutions E2S/UPPA I-SITE).

**Appendix A. Supporting information**

Supplementary data associated with this article can be found in the
online version at doi:10.1016/j.jhazmat.2025.138484.

Data availability

Data will be made available on request.

**References**

[1] Fitzgerald, W., Lamborg, C., 2005. Geochemistry of mercury. Environ Geochem 9, 107.

[2] Gilmour, C.C., Henry, E.A., 1991. Mercury methylation in aquatic systems affected by acid deposition. Environ Pollut 71 (2-4), 131-169. https://doi.org/10.1016/0269-7491(91)90031-Q.

[3] Schroeder, W.H., Munthe, J., 1998. Atmospheric mercury—an overview. Atmos Environ 32 (5), 809–822. https://doi.org/10.1016/S1352-2310(97)00293-8.

[4] Poulain, A., Amyot, M., Findlay, D., Telor, S., Barkay, T., Hintelmann, H., 2004. Biological and photochemical production of dissolved gaseous mercury in a boreal lake. Limnol Oceanogr 49 (6), 2265–2275. https://doi.org/10.4319/lo.2004.49.6.2265.

[5] Marvin-DiPasquale, M., Agee, J., McGowan, C., Oremland, R.S., Thomas, M., Krabbenhoft, D., Gilmour, C.C., 2000. Methyl-mercury degradation pathways: a comparison among three mercury-impacted ecosystems. Environ Sci Technol 34 (23), 4908–4916. https://doi.org/10.1021/es0013125.

[6] Ullrich, S.M., Tanton, T.W., Abdrashitova, S.A., 2001. Mercury in the aquatic environment: a review of factors affecting methylation. Crit Rev Environ Sci Technol 31 (3), 241–293. https://doi.org/10.1080/20016491089226.

[7] Inoko, M., 1981. Studies on the photochemical decomposition of organomercurials—methylmercury (II) chloride. Environ Pollut Ser B Chem Phys 2 (1), 3–10. https://doi.org/10.1016/0143-148X(81)90003-3.

[8] Seller, P., Kelly, C., Rudd, J., MacHutchon, A., 1996. Photodegradation of methylmercury in lakes. Nature 380 (6576), 694–697. https://doi.org/10.1038/380694a0.

[9] Chen, J., Pehkonen, S.O., Lin, C.-J., 2003. Degradation of monomethylmercury chloride by hydroxyl radicals in simulated natural waters. Water Res 37 (10), 2496–2504. https://doi.org/10.1016/S0043-1354(03)00039-3.

[10] Hammerschmidt, C.R., Fitzgerald, W.F., 2010. Iron-mediated photochemical decomposition of methylmercury in an Arctic Alaskan lake. Environ Sci Technol 44 (16), 6138–6143. https://doi.org/10.1021/es1006934.

[11] Zhang, T., Hsu-Kim, H., 2010. Photolytic degradation of methylmercury enhanced by binding to natural organic ligands. Nat Geosci 3 (7), 473–476. https://doi.org/10.1038/ngeo892.

[12] Kim, M.-K., Zoh, K.-D., 2013. Effects of natural water constituents on the photo-decomposition of methylmercury and the role of hydroxyl radical. Sci Total Environ 449, 95–101. https://doi.org/10.1016/j.scitotenv.2013.01.039.

[13] Hammerschmidt, C.R., Fitzgerald, W.F., 2006. Photodecomposition of methylmercury in an arctic Alaskan lake. Environ Sci Technol 40 (4), 1212–1216. https://doi.org/10.1021/es0513234.

[14] Li, Y., Mao, Y., Liu, G., Tachiev, G., Roelant, D., Feng, X., Cai, Y., 2010. Degradation of methylmercury and its effects on mercury distribution and cycling in the Florida Everglades. Environ Sci Technol 44 (17), 6661–6666. https://doi.org/10.1021/es1010434.

[15] Zheng, W., Hintelmann, H., 2010. Isotope fractionation of mercury during its photochemical reduction by low-molecular-weight organic compounds. J Phys Chem A 114 (12), 4246–4253. https://doi.org/10.1021/jp9111348.

[16] Bergquist, B.A., Blum, J.D., 2007. Mass-dependent and-independent fractionation of Hg isotopes by photoreduction in aquatic systems. Science 318 (5849), 417–420. https://doi.org/10.1126/science.1148050.

[17] Chandan, P., Ghosh, S., Bergquist, B.A., 2015. Mercury isotope fractionation during aqueous photoreduction of monomethylmercury in the presence of dissolved organic matter. Environ Sci Technol 49 (1), 259–267. https://doi.org/10.1021/es5034553.

[18] Rose, C.H., Ghosh, S., Blum, J.D., Bergquist, B.A., 2015. Effects of ultraviolet radiation on mercury isotope fractionation during photo-reduction for inorganic and organic mercury species. Chem Geol 405, 102–111. https://doi.org/10.1016/j.chemgeo.2015.02.025.

[19] Lotfi-Kalahroodi, E., Le Bechec, M., Tessier, E., Pigot, T., Amouroux, D., 2023. Influence of oxygen, UV light and reactive dissolved organic matter on the photodemethylation and photoreduction of monomethylmercury in model freshwater. Chemosphere 330, 138675. (https://doi.org/10.1016/j.chemosphere.2023.138675) (From NLM Medline).

[20] Sellers, P., Kelly, C.A., Rudd, J.W., 2001. Fluxes of methylmercury to the water column of a drainage lake: the relative importance of internal and external sources. Limnol Oceanogr 46 (3), 623–631. https://doi.org/10.4319/lo.2001.46.3.0623.

[21] Lehnherr, I., St. Louis, V.L., 2009. Importance of ultraviolet radiation in the photodemethylation of methylmercury in freshwater ecosystems. Environ Sci Technol 43 (15), 5692–5698. https://doi.org/10.1021/es9002923.

[22] Bouchet, S., Tessier, E., Masbou, J., Point, D., Lazzaro, X., Monperrus, M., Guédron, S., Acha, D., Amouroux, D., 2022. In situ photochemical transformation of Hg species and associated isotopic fractionation in the water column of high-altitude lakes from the Bolivian Altiplano. Environ Sci Technol 56 (4), 2258–2268. https://doi.org/10.1021/acs.est.1c04704.

[23] Amyot, M., Lean, D., Mierle, G., 1997. Photochemical formation of volatile mercury in high Arctic lakes. Environ Toxicol Chem: Int J 16 (10), 2054–2063. https://doi.org/10.1002/etc.5620161010.

[24] Black, F.J., Poulin, B.A., Flegal, A.R., 2012. Factors controlling the abiotic photo-degradation of monomethylmercury in surface waters. Geochim Et Cosmochim Acta 84, 492–507. https://doi.org/10.1016/j.gca.2012.01.019.

[25] Sun, R., Wang, D., Zhang, Y., Mao, W., Zhang, T., Ma, M., Zhang, C., 2013. Photo-degradation of monomethylmercury in the presence of chloride ion. Chemosphere 91 (11), 1471–1476. https://doi.org/10.1016/j.chemosphere.2012.12.013.

[26] O'Driscoll, N., Lean, D., Loseto, L., Carignan, R., Siciliano, S., 2004. Effect of dissolved organic carbon on the photoproduction of dissolved gaseous mercury in lakes: potential impacts of forestry. Environ Sci Technol 38 (9), 2664–2672. https://doi.org/10.1021/es034702a.

[27] Zhang, D., Yin, Y., Li, Y., Cai, Y., Liu, J., 2017. Critical role of natural organic matter in photodegradation of methylmercury in water: molecular weight and interactive effects with other environmental factors. Sci Total Environ 578, 535–541. https://doi.org/10.1016/j.scitotenv.2016.10.222.

[28] Garcia, E., Amyot, M., Ariya, P.A., 2005. Relationship between DOC photochemistry and mercury redox transformations in temperate lakes and wetlands. Geochim Et Cosmochim Acta 69 (8), 1917–1924. https://doi.org/10.1016/j.gca.2004.10.026.

[29] Qian, Y., Yin, X., Lin, H., Rao, B., Brooks, S.C., Liang, L., Gu, B., 2014. Why dissolved organic matter enhances photodegradation of methylmercury. Environ Sci Technol Lett 1 (10), 426–431. https://doi.org/10.1021/es500254z.

[30] Yang, J., Kim, J., Soerensen, A.L., Lee, W., Han, S., 2020. The role of fluorescent dissolved organic matter on mercury photoreduction rates: a case study of three temperate lakes. Geochim Et Cosmochim Acta 277, 192–205. https://doi.org/10.1016/j.gca.2020.03.027.

[31] Motta, L.C., Kritee, K., Blum, J.D., Tsz-Ki Tsui, M., Reinfelder, J.R., 2020. Mercury isotope fractionation during the photochemical reduction of Hg (II) coordinated with organic ligands. J Phys Chem A 124 (14), 2842–2853. https://doi.org/10.1021/acs.jpcja.9b06308.

[32] Zheng, W., Hintelmann, H., 2009. Mercury isotope fractionation during photoreduction in natural water is controlled by its Hg/DOC ratio. Geochim Et Cosmochim Acta 73 (22), 6704–6715. https://doi.org/10.1016/j.gca.2009.08.016.

[33] Morris, D.P., Zagarese, H., Williamson, C.E., Balseiro, E.G., Hargreaves, B.R., Modenutti, B., Moeller, R., Queimalinos, C., 1995. The attenuation of solar UV radiation in lakes and the role of dissolved organic carbon. Limnol Oceanogr 40 (8), 1381–1391. https://doi.org/10.4319/lo.1995.40.8.1381.

[34] Williamson, C.E., Stemberger, R.S., Morris, D.P., Frost, T.M., Paulsen, S.G., 1996. Ultraviolet radiation in North American lakes: attenuation estimates from DOC measurements and implications for plankton communities. Limnol Oceanogr 41 (5), 1024–1034. https://doi.org/10.4319/lo.1996.41.5.1024.

[35] Klapstein, S.J., Ziegler, S.E., Risk, D.A., O'Driscoll, N.J., 2017. Quantifying the effects of photoreactive dissolved organic matter on methylmercury photodemethylation rates in freshwaters. Environ Toxicol Chem 36 (6), 1493–1502. https://doi.org/10.1002/etc.3690.

[36] Klapstein, S.J., Ziegler, S.E., O'Driscoll, N.J., 2018. Methylmercury photodemethylation is inhibited in lakes with high dissolved organic matter. Environ Pollut 232, 392–401. https://doi.org/10.1016/j.envpol.2017.09.049.

[37] Tai, C., Li, Y., Yin, Y., Scinto, L.J., Jiang, G., Cai, Y., 2014. Methylmercury photodegradation in surface water of the Florida Everglades: importance of dissolved organic matter-methylmercury complexation. Environ Sci Technol 48 (13), 7333–7340. https://doi.org/10.1021/es500316d.
[38] Blum, J.D., Sherman, L.S., Johnson, M.W., 2014. Mercury isotopes in earth and environmental sciences. Annu Rev Earth Planet Sci 42, 249–269. https://doi.org/10.1146/annurev-earth-050212-124107.

[39] Kritee, K., Barkay, T., Blum, J.D., 2009. Mass dependent stable isotope fractionation of mercury during mer mediated microbial degradation of monomethylmercury. Geochim Et Cosmochim Acta 73 (5), 1285–1296. https://doi.org/10.1016/j.gca.2008.11.038.

[40] Kritee, K., Blum, J.D., Johnson, M.W., Bergquist, B.A., Barkay, T., 2007. Mercury stable isotope fractionation during reduction of Hg (II) to Hg (0) by mercury resistant microorganisms. Environ Sci Technol 41 (6), 1889–1895. https://doi.org/10.1021/es062019.

[41] Bergquist, B.A., Blum, J.D., 2009. The odds and evens of mercury isotopes: applications of mass-dependent and mass-independent isotope fractionation. Elements 5 (6), 353–357. https://doi.org/10.2113/gselements.5.6.353.

[42] Buchachenko, A.L., 2001. Magnetic isotope effect: Nuclear spin control of chemical reactions. J Phys Chem A 105 (44), 9995–10011.

[43] Bigeleisen, J., Mayer, M.G., 1947. Calculation of equilibrium constants for isotopic exchange reactions. J Chem Phys 15 (5), 261–267.

[44] Urey, H.C., 1947. The thermodynamic properties of isotopic substances. J Chem Soc (Resume) 562–581. https://doi.org/10.1039/JR9470000562.

[45] Bigeleisen, J., 1996. Nuclear size and shape effects in chemical reactions. Isotope chemistry of the heavy elements. J Am Chem Soc 118 (15), 3676–3680. https://doi.org/10.1021/ja954076k.

[46] Schauble, E.A., 2007. Role of nuclear volume in driving equilibrium stable isotope fractionation of mercury, thallium, and other very heavy elements. Geochim Et Cosmochim Acta 71 (9), 2170–2189. https://doi.org/10.1016/j.gca.2007.02.004.

[47] Buchachenko, A.L., Łukzen, N.N., Pedersen, J.B., 2007. On the magnetic field and isotope effects in enzymatic phosphorylation. Chem Phys Lett 434 (1-3), 139–143. https://doi.org/10.1016/j.cplett.2006.12.019.

[48] Malinovsky, D., Latruwe, K., Moens, L., Vanhaecke, F., 2010. Experimental study of mass-independence of Hg isotope fractionation during photodecomposition of dissolved methylmercury. J Anal Spectrom 25 (7), 950–956. https://doi.org/10.1039/B926650Y.

[49] Zheng, W., Hintelmann, H., 2010. Nuclear field shift effect in isotope fractionation of mercury during abiotic reduction in the absence of light. J Phys Chem A 114 (12), 4238–4245. https://doi.org/10.1021/jp910353y.

[50] Wiederhold, J.G., Cramer, C.J., Daniel, K., Infante, I., Bourdon, B., Kretzschmar, R., 2010. Equilibrium mercury isotope fractionation between dissolved Hg (II) species and thiol-bound Hg. Environ Sci Technol 44 (11), 4191–4197. DOI: https://doi.org/10.1021/es100205t.

[51] Ghosh, S., Schauble, E.A., Couloule, G.L., Blum, J.D., Bergquist, B.A., 2013. Estimation of nuclear volume dependent fractionation of mercury isotopes in equilibrium liquid-vapor evaporation experiments. Chem Geol 336, 5–12. https://doi.org/10.1016/j.chemgeol.2012.01.008.

[52] Yang, S., Liu, Y., 2016. Nuclear field shift effects on stable isotope fractionation: a review. Acta Geochim 35, 227–239. https://doi.org/10.1007/s11631-016-0109-3.

[53] Sherman, L.S., Blum, J.D., Johnson, K.P., Keeler, G.J., Barres, J.A., Douglas, T.A., 2010. Mass-independent fractionation of mercury isotopes in Arctic snow driven by sunlight. Nat Geosci 3 (3), 173–177. https://doi.org/10.1038/neo758.

[54] Jackson, T.A., Muir, D.C., Vincent, W.F., 2004. Historical variations in the stable isotope composition of mercury in Arctic lake sediments. Environ Sci Technol 38 (10), 2813–2821. https://doi.org/10.1021/es0306009.

[55] Laffont, L., Sonke, J.E., Maurice, L., Hintelmann, H., Pouilly, M., Sánchez Bacarrea, Y., Perez, T., Behra, P., 2009. Anomalous mercury isotopic compositions of fish and human hair in the Bolivian Amazon. Environ Sci Technol 43 (23), 8985–8990. https://doi.org/10.1021/es901951y.

[56] Gehrke, G.E., Blum, J.D., Slotton, D.G., Greenfield, B.K., 2011. Mercury isotopes link mercury in San Francisco Bay forage fish to surface sediments. Environ Sci Technol 45 (4), 1264–1270. https://doi.org/10.1021/es103053y.

[57] Senn, D.B., Chesney, E.J., Blum, J.D., Bank, M.S., Maage, A., Shine, J.P., 2010. Stable isotope (N, C, Hg) study of methylmercury sources and trophic transfer in the northern Gulf of Mexico. Environ Sci Technol 44 (5), 1630–1637. https://doi.org/10.1021/es902361j.

[58] Kwon, S.Y., Blum, J.D., Chen, C.Y., Meattey, D.E., Mason, R.P., 2014. Mercury isotope study of sources and exposure pathways of methylmercury in estuarine food webs in the Northeastern US. Environ Sci Technol 48 (17), 10089–10097. https://doi.org/10.1021/es5020554.

[59] Kopf, S., Ono, S., 2012. Sulfur mass-independent fractionation in liquid phase chemistry: UV photolysis of phenalkenylsulfone as a case study. Geochim Et Cosmochim Acta 85, 160–169. https://doi.org/10.1016/j.gca.2012.02.008.

[60] Malinovsky, D., Moens, L., Vanhaecke, F., 2009. Isotopic fractionation of Sn during methylation and demethylation reactions in aqueous solution. Environ Sci Technol 43 (12), 4399–4404. https://doi.org/10.1021/es9000856.

[61] Wakasa, M., Hayashi, H., Kobayashi, T., Takada, T., 1993. Enrichment of germanium-73 with the magnetic isotope effect. J Phys Chem 97 (51), 13444–13446.

[62] Gårdfeldt, K., Horvat, M., Sommar, J., Kotnik, J., Fajon, V., Wängberg, I., Lindqvist, O., 2002. Comparison of procedures for measurements of dissolved gaseous mercury in seawater performed on a Mediterranean cruise. Anal Bioanal Chem 374 (6), 1002–1008. https://doi.org/10.1007/s00216-002-1592-4.

[63] Epov, V.N., Berail, S., Jimenez-Moreno, M., Perrot, V., Pecheyran, C., Amouroux, D., Donard, O.F., 2010. Approach to measure isotopic ratios in species using multicollector-ICPMS coupled with chromatography. Anal Chem 82 (13), 5652–5662. https://doi.org/10.1021/ac100648f.

[64] Bérail, S., Cavalheiro, J., Tessier, E., Barre, J.P., Pedrero, Z., Donard, O.F.X., Amouroux, D., 2017. Determination of total Hg isotopic composition at ultra-trace

levels by on line cold vapor generation and dual gold-amalgamation coupled to MC-ICP-MS. J Anal Spectrom 32 (2), 373–384. https://doi.org/10.1039/C6JA00375C.

[65] Jiménez-Moreno, M., Perrot, V., Epov, V.N., Monperrus, M., Amouroux, D., 2013. Chemical kinetic isotope fractionation of mercury during abiotic methylation of Hg (II) by methylcobalamin in aqueous chloride media. Chem Geol 336, 26–36. https://doi.org/10.1016/j.chemgeol.2012.08.029.

[66] Blum, J.D., Bergquist, B.A., 2007. Reporting of variations in the natural isotopic composition of mercury. Anal Bioanal Chem 388 (2), 353–359. https://doi.org/10.1007/s00216-007-1236-9.

[67] Ronzani, F., Arzoumanian, E., Blanc, S., Bordat, P., Pigot, T., Cugnet, C., Oliveros, E., Sarakha, M., Richard, C., Lacombe, S., 2013. Efficient cyanoaromatic photosensitizers for singlet oxygen production: synthesis and characterization of the transient reactive species. Phys Chem Chem Phys 15 (40), 17219–17232. DOI: DOI:10.1039/c3cp52168k.

[68] Zhao, J., Nelson, D.J., 2005. Fluorescence study of the interaction of Suwannee River fulvic acid with metal ions and Al3+-metal ion competition. J Inorg Biochem 99 (2), 383–396. https://doi.org/10.1016/j.jinorgbio.2004.10.005.

[69] Kim, D., Grassian, V.H., 2021. Attenuated total reflection-fourier transform infrared and atomic force microscopy-infrared spectroscopic investigation of suwannee river Fulvic acid and its interactions with α-FeOOH. ACS Earth Space Science 6 (1), 81–89. https://doi.org/10.1021/acsearthspacechem.1c00288.

[70] Yoon, T.H., Johnson, S.B., Brown, G.E., 2004. Adsorption of Suwannee River fulvic acid on aluminum oxyhydroxide surfaces: an in situ ATR-FITIR study. Langmuir 20 (14), 5655–5658. https://doi.org/10.1021/la0499214.

[71] Larrivee, E.M., Elkins, K.M., Andrews, S.E., Nelson, D.J., 2003. Fluorescence characterization of the interaction of Al3+ and Pd2+ with Suwannee River fulvic acid in the absence and presence of the herbicide 2, 4-dichlorophenoxyacetic acid. J Inorg Biochem 97 (1), 32–45. https://doi.org/10.1016/S0162-0134(03)00239-3.

[72] Cottrell, B.A., Timko, S.A., Devera, L., Robinson, A.K., Gonsior, M., Vizenor, A.E., Simpson, A.J., Cooper, W.J., 2013. Photochemistry of excited-state species in natural waters: a role for particulate organic matter. Water Res 47 (14), 5189–5199. https://doi.org/10.1016/j.watres.2013.05.059.

[73] Duval, B., Tessier, E., Kortazar, L., Fernández, L.A., de Diego, A., Amouroux, D., 2023. Dynamics, distribution, and transformations of mercury species from pyrenean high-altitude lakes. Environ Res 216, 114611. https://doi.org/10.1012/acs.est.1c04704.

[74] Monperrus, M., Tessier, E., Veschambre, S., Amouroux, D., Donard, O., 2005. Simultaneous speciation of mercury and butyltin compounds in natural waters and snow by propylation and species-specific isotope dilution mass spectrometry analysis. Anal Bioanal Chem 381 (4), 854–862. https://doi.org/10.1007/s00216-004-2973-7.

[75] Clémens, S., Monperrus, M., Donard, O.F., Amouroux, D., Guérin, T., 2011. Mercury speciation analysis in seafood by species-specific isotope dilution: method validation and occurrence data. Anal Bioanal Chem 401 (9), 2699–2711. https://doi.org/10.1007/s00216-011-5040-1.

[76] Martin-Doimeadios, R.R.G., Tessier, E., Amouroux, D., Guyoneaud, R., Duran, R., Caumette, P., Donard, O., 2004. Mercury methylation/demethylation and volatilization pathways in estuarine sediment slurries using species-specific enriched stable isotopes. Mar Chem 90 (1-4), 107–123. https://doi.org/10.1016/j.marchem.2004.02.022.

[77] Martin-Doimeadios, R.R., Monperrus, M., Krupp, E., Amouroux, D., Donard, O., 2003. Using speciated isotope dilution with GC–inductively coupled plasma MS to determine and unravel the artificial formation of monomethylmercury in certified reference sediments. Anal Chem 75 (13), 3202–3211. https://doi.org/10.1021/ac026411a.

[78] Enrico, M., Mere, A., Zhou, H., Carrier, H., Tessier, E., Bouyssiere, B., 2019. Experimental tests of natural gas samplers prior to mercury concentration analysis. Energy Fuels 34 (5), 5205–5212. https://doi.org/10.1021/acs.energyfuels.9b03540.

[79] Blum, J.D., Johnson, M.W., 2017. Recent developments in mercury stable isotope analysis. Rev Mineral Geochem 82 (1), 733–757. https://doi.org/10.2138/rmg.2017.82.17.

[80] Mariotti, A., Germon, J., Hubert, P., Kaiser, P., Letolle, R., Tardieux, A., Tardieux, P., 1981. Experimental determination of nitrogen kinetic isotope fractionation: some principles; illustration for the denitrification and nitrification processes. Plant Soil 62, 413–430. https://doi.org/10.1007/BF02374138.

[81] York, D., Evensen, N.M., Martínez, M.L., De Basabe Delgado, J., 2004. Unified equations for the slope, intercept, and standard errors of the best straight line. Am J Phys 72 (3), 367–375. https://doi.org/10.1119/1.1632486.

[82] Janssen, S.E., Lepak, R.F., Tate, M.T., Ogorek, J.M., DeWild, J.F., Babiarz, C.L., Hurley, J.P., Krabbenhoft, D.P., 2019. Rapid pre-concentration of mercury in solids and water for isotopic analysis. Anal Chim Acta 1054, 95–103. https://doi.org/10.1016/j.aca.2018.12.026.

[83] Buchachenko, A.L., 2018. Magnetic isotopes as a means to elucidate Earth and environmental chemistry. Russ Chem Rev 87 (8), 727. https://doi.org/10.1070/RCR4818.

[84] Turro, N.J., 1983. Influence of nuclear spin on chemical reactions: magnetic isotope and magnetic field effects (a review). Proc Natl Acad Sci 80 (2), 609–621. https://doi.org/10.1073/pnas.80.2.609.

[85] Epov, V.N., 2011. Magnetic isotope effect and theory of atomic orbital hybridization to predict a mechanism of chemical exchange reactions. Phys Chem Chem Phys 13 (29), 13222–13231. https://doi.org/10.1039/C1CP21012B.

[86] Turro, N.J., Ramamurthy, V., Scaiano, J.C., 2009. Principles of molecular photochemistry: an introduction. University Science Books.
[87] Mosinger, J., Lang, K., Plistil, L., Jesenská, S., Hostomský, J., Zelinger, Z., Kubát, P., 2010. Fluorescent polyurethane nanofabrics: a source of singlet oxygen and oxygen sensing. Langmuir 26 (12), 10050–10056, 10.1021/la1001607.

[88] Zhang, H., 2006. Photochemical redox reactions of mercury. Recent Dev Mercury Sci 37–79.

[89] Niu, X.Z., Liu, C., Gutierrez, L., Croué, J.P., 2014. Photobleaching-induced changes in photosensitizing properties of dissolved organic matter. Water Res 66, 140–148. <https://doi.org/10.1016/j.watres.2014.08.017> From NLM.

[90] Helms, J.R., Stubbins, A., Ritchie, J.D., Minor, E.C., Kieber, D.J., Mopper, K., 2008. Absorption spectral slopes and slope ratios as indicators of molecular weight, source, and photobleaching of chromophoric dissolved organic matter. Limnol Oceanogr 53 (3), 955–969. https://doi.org/10.4319/lo.2008.53.3.0955.

[91] Buchachenko, A.L., 2013. Mass-independent isotope effects. J Phys Chem B 117 (8), 2231–2238. https://doi.org/10.1021/jp308727w.

[92] Motta, L.C., Chien, A.D., Rask, A.E., Zimmerman, P.M., 2020. Mercury magnetic isotope effect: a plausible photochemical mechanism. J Phys Chem A 124 (19), 3711–3719. https://doi.org/10.1021/acs.jpca.0c00661.

[93] Fernández-Gómez, C., Drott, A., Bjorn, E., Díez, S., Bayona, J.M., Tesfalidet, S., Lindfors, A., Skyllberg, U., 2013. Towards universal wavelength-specific photodegradation rate constants for methyl mercury in humic waters, exemplified by a boreal lake-wetland gradient. Environ Sci Technol 47 (12), 6279–6287. https://doi.org/10.1021/es400373s.

[94] Zhang, X., Li, Y., Feng, G., Tai, C., Yin, Y., Cai, Y., Liu, J., 2018. Probing the DOM-mediated photodegradation of methylmercury by using organic ligands with different molecular structures as the DOM model. Water Res 138, 264–271. https://doi.org/10.1016/j.watres.2018.03.055.