# Record of trace organic contaminants in a river sediment core: A fingerprint of regulatory changes?

Thomas Thiebault, Pierre Labadie, Anthony Foucher, Jean-Sébastien Barbier, Fabrice Alliot, Thibaut Soubise, Patrick Pardon, Hélène Budzinski, Elie Dhivert, Olivier Evrard, et al.

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## To cite this version:

Thomas Thiebault, Pierre Labadie, Anthony Foucher, Jean-Sébastien Barbier, Fabrice Alliot, et al. Record of trace organic contaminants in a river sediment core: A fingerprint of regulatory changes?. Journal of Environmental Management, 2025, 389, pp.126276. 10.1016/j.jenvman.2025.126276. hal-05119634

HAL Id: hal-05119634

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Submitted on 18 Jun 2025

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Research article

# Record of trace organic contaminants in a river sediment core: A fingerprint of regulatory changes?

Thomas Thiebault<sup>a, *</sup>, Pierre Labadie<sup>b</sup>, Anthony Foucher<sup>c</sup>, Jean-Sébastien Barbier<sup>c</sup>, Fabrice Alliot<sup>a</sup>, Thibaut Soubise<sup>b, d</sup>, Patrick Pardon<sup>b</sup>, Hélène Budzinski<sup>b</sup>, Elie Dhivert<sup>d</sup>, Olivier Evrard<sup>c, d</sup>, Sophie Ayrault<sup>c, d</sup>

## ARTICLE INFO

Keywords:

Sediments

Wastewater treatment

Trace organic contaminants

Pharmaceuticals

Pesticides

PFAS

## ABSTRACT

This study reconstructs the historical deposition of 76 trace organic contaminants (TrOCs), comprising pharmaceuticals, pesticides, and per- and polyfluoroalkyl substances (PFAS), in a dated sediment core taken from a secondary channel of the Seine River downstream of the Paris conurbation. Sediment chronology, established through ²¹⁰Pb and ¹³⁷Cs radionuclide depth activities, spans the period from 1943 to 2022, providing a long-term record of anthropogenic influence on fluvial contamination. The vertical profiles of pharmaceutical residues reveal temporal patterns closely aligned with their respective dates of market introduction, followed by increasing concentrations and subsequent declines. Maximum TrOC sedimentary concentrations were observed during the 1980s, coinciding with incomplete wastewater collection coverage. A sharp decrease in concentrations post-2000 reflects significant improvements in wastewater collection and treatment infrastructure. For pesticides, sedimentary trends also correlate with usage history and regulatory actions, with detectable declines frequently preceding official bans, suggesting the influence of progressive restrictions or shifts in agricultural practices. In contrast, PFAS profiles display more complex dynamics: legacy compounds such as perfluorooctanesulfonic acid and its precursors peaked in the early 2000s, after which their decline was offset by the increasing presence of alternative substances (e.g., fluorotelomersulfonic acids, betaines), indicative of industrial substitution. Overall, sediment composition had a secondary influence on TrOC profiles, with external factors such as regulatory decisions and technological developments in wastewater treatment playing predominant roles. These findings underscore the utility of collecting and analyzing sedimentary archives for assessing the long-term effectiveness of environmental policies and for identifying persistent and emerging contaminants in riverine systems.

## 1. Introduction

Since the WWII, many trace organic contaminants (TrOCs) have been regularly introduced on the market for different uses (Bernhardt et al., 2017), e.g., pharmaceutical products, per and polyfluoroalkyl compounds (i.e., PFAS) and pesticides. These molecules are generally classified as contaminants of emerging concern, meaning that their quantification in environmental compartments is relatively recent and raises significant concerns about their potential impact on human and ecosystem health (Daughton, 2004).

Moreover, as no environmental impact evaluation was performed prior to market launch (until a pesticide directive by the European Commission (1991a)), a significant and growing number of TrOCs (as main products or metabolites) have been detected in numerous environmental compartments, for example, since the 1960s and 1970s for some pesticides (Robeck et al., 1965) and pharmaceutical products (Hignite and Azarnoff, 1977) and in the late 1990s for PFAS (Gaines, 2023).

In the European Union, two political decisions had a major effect on raising awareness about the quality of water bodies: first, the Urban

<footnote><sup>a</sup> EPHE, PSL Univ., Sorbonne Univ., CNRS, UMR 7619 METIS, F-75005, Paris, France</footnote>

<footnote><sup>b</sup> Univ. Bordeaux, CNRS, Bordeaux INP, EPOC, UMR 5805, F-33600, Pessac, France</footnote>

<footnote><sup>c</sup> Laboratoire des Sciences du Climat et de l'Environnement (LSCE-IPSL), Université Paris-Saclay, UMR 8212 (CEA-CNRS-UVSQ), F-91190, Gif-sur-Yvette, France</footnote>

<footnote><sup>d</sup> Université de Lyon, Université Claude Bernard Lyon 1, CNRS, ENTPE, UMR 5023 LEHNA, F-69518, Vaulx-en-Velin, France</footnote>

<footnote>* Corresponding author.</footnote>

<footnote>E-mail address: thomas.thiebault@ephe.psl.eu (T. Thiebault).</footnote>
Wastewater treatment Directive (UWD) in 1991 (European Commission, 1991b), and then, in 2000, the Water Framework Directive (European Commission, 2000). These two European directives aimed to improve the quality of water environments in Europe by targeting the collection and treatment of wastewater and ensuring the good chemical and ecological status of water bodies in the future. These political decisions were accompanied by obligations to monitor an ever-increasing number of TrOCs in natural environments (Loos et al., 2009; McCance et al., 2018). However, depending on the targeted TrOCs, the contamination of environmental compartments can be much older than the first detection, hindering the proper understanding of the contamination trajectory of water bodies.

Sedimentary archives are considered particularly relevant for reconstructing this historical contamination, which has already been demonstrated for hydrophobic TrOCs such as PAHs or PCBs (Ayrault et al., 2020; Dendievel et al., 2020). Furthermore, recent studies have demonstrated that more hydrophilic and amphiphilic contaminants, such as pesticides (excluding organochlorine pesticides such as DDT) (Chiaia-Hernández et al., 2020), PFAS (Mussabek et al., 2019) and pharmaceutical products, may also be recorded in sedimentary archives (Lahti and Oikari, 2012), revealing the relevance of using sedimentary archives for reconstructing the temporal records of TrOC deposition.

However, once the vertical distribution of contaminants has been obtained, interpreting the patterns may be challenging (Soriano et al., 2024). Generally, most authors concur that the oldest detection of a specific contaminant is concurrent with or posterior to the market launch year of the investigated contaminant (Klaminder et al., 2015). Another strong assumption is that the sedimentary occurrence of contaminants is related to their dissolved concentration and solid/water partition coefficient (Lara-Martín et al., 2015; Thiebault et al., 2017). However, unlike hydrophobic contaminants, for which sedimentary occurrences are considered proportional to historical environmental contamination based on the organic fraction of sediment (Castro and Vale, 1995), for more hydrophilic contaminants, variations in both hydrosedimentary conditions (da Silva et al., 2011; Zhou and Broodbank, 2014) and sedimentary compositions (Al-Khazraji and Boxall, 2016; Kodešová et al., 2015) may impact solid/water partition non-linearly. It is therefore important to assess the vertical variation in sediment composition (granulometry, organic content) to avoid any potential misinterpretation of the observed pattern. The correct understanding of the vertical distribution of TrOCs is also dependent on the development of an appropriate age model to date such an archive (Fuller et al., 1999).

The vertical variations in the concentration of rather hydrophilic contaminants therefore depend on several factors, such as (i) the adsorption potential of the sediment; (ii) the amount of contaminant used in the watershed; and (iii) the impact of potential filters between the contaminant source and the watercourse. With respect to the latter point, a proper evaluation of the main sources of the contaminants investigated is very important. For example, it has been known for some time that most pharmaceutical products (with the exception of livestock production sources) are discharged into the environment through wastewater treatment plants (WWTPs), which do not fully remove TrOCs (Tran et al., 2018). However, both sewer networks and wastewater treatment technologies have evolved considerably over the last 60 years (De Feo et al., 2014). Pesticides discharged into the environment are more difficult to trace, as even though most pesticides are used in the field for crop protection, generating a diffuse source toward water bodies, a significant amount of pesticides are also rejected by point sources, such as wastewater and stormwater discharge points (Botta et al., 2009; Zgheib et al., 2012).

Moreover, the vertical distribution of TrOCs such as pharmaceutical products, PFAS and pesticides within sedimentary archives deserves a careful investigation into the historical evolution of the watershed is required to interpret correctly the reconstructed trends. Therefore, in this study, we investigated the occurrence of 70 TrOCs in a sedimentary

archive collected downstream of Paris conurbation in the Seine River. The targeted TrOCs were pharmaceutical products, PFAS and pesticides, including both parent compounds and some major metabolites. From an analysis of their vertical distribution, the objectives were (i) to assess a long-term chronology of TrOC contamination highlighting the contamination trajectories since WWII; (ii) to evaluate the impacts of the hydrosedimentary conditions and associated sedimentary composition; (iii) to evaluate the impacts of WWTP evolution and change in substance use in the watershed on TrOC sedimentary occurrence; and (iv) to assess the impact of use restrictions on sedimentary contamination reflecting the overall environmental quality.

## 2. Materials and methods

### 2.1. Site settings and coring operations

The investigated site is an abandoned channel of the Seine River, the "bras de ma campagne", located in Vernon 55 km downstream from Paris on the Seine River (Fig. 1). This coring station was selected after analyzing aerial images and historical maps of the region, and because the Seine River Basin is particularly impacted by anthropogenic activities. It is inhabited by about 30 % of the French population (including Paris megacity), and it concentrates 40 % of French industrial activities and nearly 25 % of the national agricultural activity. Therefore, this site was considered to be particularly relevant for the analysis of historical contamination by TrOCs of various classes.

It corresponds to an old secondary channel still connected from its downstream mouth in its present form and inherited for fluvial islands networks in the mid-19th century (Fig. 1). A longitudinal transect was made in this depositional environment and three cores were taken in March 2022 from the confluence with the Seine downstream to the most upstream part of the site under water (the staff map shows downstream and upstream connections, but the upstream connection is now closed; Fig. 4B). The cores (from 48 to 83 cm long) were taken using the LSCE coring platform, which was equipped with a 63 mm diameter UWITEC gravity corer. Preliminary age models were constructed for these three archives. On the one hand, the reconstructed ages led to exclude the upstream and downstream cores, which were too affected by the occurrence of instantaneous events (e.g., floods) that affected the ¹³⁷Cs and ²¹⁰Pbₓ profiles. On the other hand, the core located in the central part of the abandoned secondary channel shows good consistency between the profiles obtained for these two radioisotopes. In addition, it recorded a large number of coarser deposits associated with historical floods, which is particularly useful for validating the age model.

The final core sampling for TrOC analysis was conducted in July 2022 using a UWITEC gravity corer with a large diameter (i.e., 90 mm). This campaign resulted in the collection of a 2.06 m long core in the central part of the secondary abandoned channel referred to as VRN22-C02 (IGSN number 10.58052/IEFOU002G).

Immediately after coring, the core was aseptically cut in half lengthwise in a confined chamber. One half of the core was regularly subsampled (1 cm taken for every 5 cm of sediment) and was used for TrOC geochemical analyses on freeze-dried bulk samples and the other half was identically subsampled for sediment dating (storage at 4 °C).

### 2.2. Sediment analysis and dating

The particle-size was determined by an enhanced laser diffraction particle size analyzer (Beckman Coulter, LS230). The total organic carbon (TOC) was determined by using Rock-Eval pyrolysis.

The analyses of artificial radionuclides (¹³⁷Cs, ²⁴¹Am and ²¹⁰Pbₓ isotopes) on 33 samples collected along the sediment core were performed using coaxial HPGe detectors (HTDS/Mirion). The age-depth model was constructed using the SERAC package computed with R (Bruel and Sabatier, 2020)), by applying the CF:CS model (Constant Flux: Constant Sedimentation; Appleby and Oldfield (1978)) after
<figure><img src="image_3.png" /><figcaption>**Fig. 1.** Location map of the Vernon coring site at the "Bras de ma campagne", which is based on (A) the general staff map in the mid-19th century and (B) a current aerial photograph (source: www.geoportail.gouv.fr).</figcaption></figure>

identifying and removing instantaneous deposits (e.g., floods) for calculation. The $^{210}$Pb$_{ex}$ age model was validated by comparing the output of the model with the first detection of the $^{137}$Cs time marker in 1954 and with the identification of peaks attributable to the global atmospheric thermonuclear weapon tests in 1963 and to the Chernobyl fallout in 1986. The $^{241}$Am, deposited in France during the thermonuclear weapon tests only, allowed to identify unambiguously the origin of the two $^{137}$Cs peaks.

Hereafter, samples are labeled according to the sediment level or the year of the sediment layer deposition.

## 2.3. Sample preparation and TrOC analysis

In this study, 76 molecules were targeted and divided into three classes, (i) 26 pharmaceutical products covering different therapeutic classes, (ii) 16 pesticides, with a dominance of herbicides and two fungicides, and (iii) 34 PFAS, which were divided into subgroups according to their chemical structure with PFCAs (perfluoroalkyl carboxylic acids, 10 compounds), PFSAs (perfluoroalkyl sulfonic acids, 4 compounds), PFOS (perfluorooctanesulfonic acid), PFOS Precursors (12 compounds) Betaines (2 compounds) and FTSAs (fluorotelomersulfonic acids, 4 compounds). The sample extraction and quantification procedures were adapted from previous studies carried out for pharmaceuticals and pesticides (Thiebault et al., 2021) and for PFAS (Macorps et al., 2023). Briefly, sediments (1 g dry weight, d.w.) were supplemented with internal standards (1 ng each for PFAS, 50 ng for pesticides and pharmaceuticals) and extracted via i) accelerated solvent extraction (ASE350, ThermoFisher Scientific, Illkirch-Graffenstaden, France) with methanol (MeOH) as the solvent, and the operating temperature and pressure were 100 °C and 1000 psi, respectively, for pharmaceuticals and pesticides and ii) microwaved-assisted solvent extraction with 250 mM ammonium acetate (CH$_3$COONH$_4$) in MeOH (12 mL0), followed by filtration on glass wool and graphitized carbon clean-up for PFAS. The extracts were evaporated to approximately 300 µL under a gentle nitrogen stream at 40 °C and stored at -20 °C prior to analysis. Targeted PFAS analysis was performed via liquid chromatography electrospray ionization coupled with tandem mass spectrometry (LC-ESI-MS/MS) on a 1290 LC system interfaced with a 6495 triple quadrupole mass spectrometer (Agilent Technologies, Massy, France), whereas targeted pesticide and pharmaceutical analysis was performed via a 1260 LC system followed by a 6465 triple quadrupole mass spectrometer.

LC-MS-grade acetonitrile (AcN) and MeOH were purchased from VWR, and ultrapure water (H$_2$O) was prepared with a Pure Lab Chorus water purification system (Veolia Water). Procedural blanks were also performed for validation. The general properties of the investigated TrOCs are presented in Table 1, and full quantification results are

available in Table S1.

## 2.4. Validation

The limits of quantification (LOQ) and limits of detection (LOD) of the method were calculated via the signal-to-noise (S/N) ratios. The LOD and LOQ were determined by extracting spiked (100 ng g$^{-1}$ for pesticides and pharmaceuticals, 5 ng g$^{-1}$ for PFAS) matrix-matched (n = 5) sediment samples. The LOD and LOQ were subsequently calculated on the basis of the ratio between the spiked amount and the S/N ratio multiplied by 3 and 10 for the LOD and LOQ, respectively. The recovery ratios were also evaluated in these experiments (similar spiking concentration, n = 5) via the same extraction and quantification protocol as those used for the samples. Finally, during each analytical run, blanks and quantification controls were regularly used to crosscheck the quantification quality. Detailed validation parameters are provided in recent publications (Macorps et al., 2023; Thiebault et al., 2021), and LOD are given in Table S1.

# 3. Results and discussion

## 3.1. Sediment characterization

The VRN22C02 core is 206 cm in length. The upper unit (0–56 cm) was light brown with a silty texture and a density between 0.8 and 1.0 g cm$^{-3}$. A few millimeter-sized organic residues were visible. The second lower unit (56–206 cm) was brown-black in color, with a distinct odor of hydrocarbons. The density was higher, between 1.0 and 1.3 g cm$^{-3}$ and several sandy streaks were visible, corresponding to flood levels. These flood levels were confirmed by granulometric analysis (presence of coarser material) and by tomodensitometric analysis (presence of hyperdense bands, Fig. 2). In addition, thin hyperdense layers that were not associated with coarser deposits were identified in this sedimentary archive (n = 21). A total of 28 events were observed, with corresponding thicknesses ranging from 2 to 25 mm.

The granulometry of the core was fairly stable along the sedimentary sequence, as it contained between 30 and 40 % of sand, 60–70 % of silt and a maximum of 3 % of clay. The material was therefore predominantly silty to sandy-sized. There were a few sandier levels (n = 7) corresponding to flood levels (Fig. 2). The total organic carbon content in these sediment layers varied between 15.2 % and 3.2 %, with a general tendency for the TOC content to increase with depth. This transition in TOC content suggests a decrease on the vegetation of the channel, indicating an increase in water flow, resulting from a hydrological connection at this level (i.e., 140 cm) between the Seine mainstream and the cored secondary abandoned channel, which appeared to
<caption>**Table 1**</caption>

General properties of the targeted TrOCs with Ab. the abbreviation, M<sub>w</sub> is the molecular weight in g.mol<sup>-1</sup>, log Kow the octanol/water partition coefficient, Charge the charge state at pH = 7 and MAD the market authorization date, * indicates byproducts and § for PFAS with separate quantification of linear and branched forms. Data extracted from Chempider Database (2025) unless otherwise stated.

<table><thead><tr><th>Type</th><th>Name</th><th>Ab.</th><th>Formula</th><th>log K<sub>ow</sub></th><th>Charge</th><th>MAD<sup>a,b</sup></th></tr></thead><tbody><tr><td colspan="7"><strong>Pharmaceutical Product</strong></td></tr><tr><td rowspan="12">Antibiotics</td><td>Amoxicillin</td><td>AMO</td><td>C<sub>16</sub>H<sub>19</sub>N<sub>3</sub>O<sub>5</sub>S</td><td>0.97</td><td>+/-</td><td>1972</td></tr><tr><td>Tetracycline</td><td>TET</td><td>C<sub>22</sub>H<sub>24</sub>N<sub>2</sub>O<sub>8</sub></td><td>-1.33</td><td>+</td><td>1949</td></tr><tr><td>Trimethoprim</td><td>TMP</td><td>C<sub>14</sub>H<sub>18</sub>N<sub>4</sub>O<sub>3</sub></td><td>0.73</td><td>0/+</td><td>1968</td></tr><tr><td>Ciprofloxacin</td><td>CIP</td><td>C<sub>17</sub>H<sub>18</sub>FN<sub>3</sub>O<sub>3</sub></td><td>0.28</td><td>+/-</td><td>1987</td></tr><tr><td>Enrofloxacin</td><td>ENR</td><td>C<sub>19</sub>H<sub>22</sub>FN<sub>3</sub>O<sub>3</sub></td><td>0.70</td><td>-</td><td>1970's</td></tr><tr><td>Norfloxacin</td><td>NOR</td><td>C<sub>16</sub>H<sub>18</sub>FN<sub>3</sub>O<sub>3</sub></td><td>-0.31</td><td>+/-</td><td>1985</td></tr><tr><td>Ofloxacin</td><td>OFL</td><td>C<sub>18</sub>H<sub>20</sub>FN<sub>3</sub>O<sub>4</sub></td><td>-2.00</td><td>-</td><td>1986</td></tr><tr><td>Erythromycin</td><td>ERY</td><td>C<sub>37</sub>H<sub>67</sub>NO<sub>13</sub></td><td>3.06</td><td>+</td><td>1952</td></tr><tr><td>Flumequine</td><td>FLQ</td><td>C<sub>14</sub>H<sub>12</sub>FNO<sub>3</sub></td><td>2.60</td><td>-</td><td>1976</td></tr><tr><td>Oxolinic Acid</td><td>OXO</td><td>C<sub>13</sub>H<sub>11</sub>NO<sub>5</sub></td><td>1.70</td><td>-</td><td>1974</td></tr><tr><td>Sulfamethoxazole</td><td>SMX</td><td>C<sub>10</sub>H<sub>11</sub>N<sub>3</sub>O<sub>3</sub>S</td><td>0.89</td><td>-</td><td>1968</td></tr><tr><td>Sulfamethazine</td><td>SMZ</td><td>C<sub>12</sub>H<sub>14</sub>N<sub>4</sub>O<sub>2</sub>S</td><td>0.89</td><td>0/-</td><td>1949</td></tr><tr><td rowspan="6">Analgesics</td><td>Acetaminophen</td><td>ACM</td><td>C<sub>8</sub>H<sub>9</sub>NO<sub>2</sub></td><td>0.46</td><td>0</td><td>1957</td></tr><tr><td>Codeine</td><td>COD</td><td>C<sub>18</sub>H<sub>21</sub>NO<sub>3</sub></td><td>1.28</td><td>+</td><td>1840s</td></tr><tr><td>Tramadol</td><td>TRA</td><td>C<sub>16</sub>H<sub>25</sub>NO<sub>2</sub></td><td>3.01</td><td>+</td><td>1997</td></tr><tr><td>Carbamazepine</td><td>CBZ</td><td>C<sub>15</sub>H<sub>12</sub>N<sub>2</sub>O</td><td>2.45</td><td>0</td><td>1974</td></tr><tr><td>β-blockers</td><td>Atenolol</td><td>ATE</td><td>C<sub>14</sub>H<sub>22</sub>N<sub>2</sub>O<sub>3</sub></td><td>0.16</td><td>+</td><td>1978</td></tr><tr><td>Metoprolol</td><td>MET</td><td>C<sub>15</sub>H<sub>25</sub>NO<sub>3</sub></td><td>1.88</td><td>+</td><td>1974</td></tr><tr><td rowspan="2">Hypolipidemic</td><td>Propranolol</td><td>PRP</td><td>C<sub>16</sub>H<sub>21</sub>NO<sub>2</sub></td><td>2.60</td><td>+</td><td>1965</td></tr><tr><td>Bezafibrate</td><td>BZB</td><td>C<sub>16</sub>H<sub>20</sub>ClNO<sub>4</sub></td><td>4.25</td><td>-</td><td>1982</td></tr><tr><td>Non-Steroidal</td><td>Diclofenac</td><td>DCF</td><td>C<sub>14</sub>H<sub>11</sub>Cl<sub>2</sub>NO<sub>2</sub></td><td>4.02</td><td>-</td><td>1950's</td></tr><tr><td>Anti-Inflamm.</td><td>Ketoprofen</td><td>KET</td><td>C<sub>16</sub>H<sub>14</sub>O<sub>3</sub></td><td>3.00</td><td>-</td><td>1970's</td></tr><tr><td rowspan="4">Psychotropic drugs</td><td>Diazepam</td><td>DIA</td><td>C<sub>16</sub>H<sub>13</sub>ClN<sub>2</sub>O</td><td>2.70</td><td>0</td><td>1963</td></tr><tr><td>Oxazepam</td><td>OXA</td><td>C<sub>15</sub>H<sub>11</sub>ClN<sub>2</sub>O<sub>2</sub></td><td>3.34</td><td>0</td><td>1965</td></tr><tr><td>Fluoxetine</td><td>FLX</td><td>C<sub>17</sub>H<sub>18</sub>F<sub>3</sub>NO</td><td>4.65</td><td>+</td><td>1989</td></tr><tr><td>Venlafaxine</td><td>VEN</td><td>C<sub>17</sub>H<sub>27</sub>NO<sub>2</sub></td><td>3.28</td><td>+</td><td>1993</td></tr><tr><td colspan="7"><strong>Pesticides</strong></td></tr><tr><td rowspan="2">Fungicides</td><td>Carbendazim</td><td>CBD</td><td>C<sub>9</sub>H<sub>9</sub>N<sub>3</sub>O<sub>2</sub></td><td>1.55</td><td>0</td><td>1967</td></tr><tr><td>Tebucazole</td><td>TEB</td><td>C<sub>16</sub>H<sub>22</sub>ClN<sub>3</sub>O</td><td>3.89</td><td>0</td><td>1986</td></tr><tr><td rowspan="12">Herbicides</td><td>Metazachlor</td><td>MTZ</td><td>C<sub>14</sub>H<sub>16</sub>ClN<sub>3</sub>O</td><td>2.38</td><td>0</td><td>1983</td></tr><tr><td>Metolachlor</td><td>MTC</td><td>C<sub>15</sub>H<sub>22</sub>ClNO<sub>2</sub></td><td>3.24</td><td>0</td><td>1975</td></tr><tr><td>Oxadiazon</td><td>OXD</td><td>C<sub>15</sub>H<sub>18</sub>Cl<sub>2</sub>N<sub>2</sub>O<sub>3</sub></td><td>5.33</td><td>-</td><td>1969</td></tr><tr><td>Diflufenican</td><td>DFF</td><td>C<sub>19</sub>H<sub>11</sub>F<sub>5</sub>N<sub>2</sub>O<sub>2</sub></td><td>4.2</td><td>0</td><td>1988</td></tr><tr><td>3,4 Dichloroaniline*</td><td>DCA</td><td>C<sub>6</sub>H<sub>5</sub>Cl<sub>2</sub>N</td><td>2.37</td><td>0</td><td>-</td></tr><tr><td>Chlortoluron</td><td>CHT</td><td>C<sub>10</sub>H<sub>13</sub>ClN<sub>2</sub>O</td><td>2.58</td><td>0</td><td>1970</td></tr><tr><td>Isoproturon</td><td>ISO</td><td>C<sub>12</sub>H<sub>18</sub>N<sub>2</sub>O</td><td>2.84</td><td>0</td><td>1974</td></tr><tr><td>Nicosulfuron</td><td>NIC</td><td>C<sub>15</sub>H<sub>18</sub>N<sub>6</sub>O<sub>6</sub>S</td><td>0.6</td><td>-</td><td>1993</td></tr><tr><td>Prosulfuron</td><td>PRS</td><td>C<sub>15</sub>H<sub>16</sub>F<sub>3</sub>N<sub>5</sub>O<sub>4</sub>S</td><td>1.50</td><td>-</td><td>1999</td></tr><tr><td>Atrazine</td><td>ATZ</td><td>C<sub>6</sub>H<sub>14</sub>ClN<sub>5</sub></td><td>2.82</td><td>0</td><td>1959</td></tr><tr><td>Desethylatrazine*</td><td>DEA</td><td>C<sub>6</sub>H<sub>10</sub>ClN<sub>5</sub></td><td>1.51</td><td>0</td><td>-</td></tr><tr><td>Deisopropylatrazine*</td><td>DIA</td><td>C<sub>6</sub>H<sub>8</sub>ClN<sub>5</sub></td><td>1.15</td><td>0</td><td>-</td></tr><tr><td>Hydroxyatrazine*</td><td>HTZ</td><td>C<sub>6</sub>H<sub>15</sub>N<sub>5</sub>O</td><td>2.09</td><td>0</td><td>-</td></tr><tr><td>Simazine</td><td>SIM</td><td>C<sub>7</sub>H<sub>12</sub>ClN<sub>5</sub></td><td>2.40</td><td>0</td><td>1957</td></tr><tr><td colspan="7"><strong>PFAS</strong></td></tr><tr><td rowspan="12">PFCAs</td><td>Perfluoropentanoic acid</td><td>PFPeA</td><td>C<sub>8</sub>HF<sub>9</sub>O<sub>2</sub></td><td>5.29</td><td>-</td><td>n.d.</td></tr><tr><td>Perfluorohexanoic acid</td><td>PFHxA</td><td>C<sub>6</sub>HF<sub>11</sub>O<sub>2</sub></td><td>5.79</td><td>-</td><td>n.d.</td></tr><tr><td>Perfluoroheptanoic acid</td><td>PFHpA</td><td>C<sub>7</sub>HF<sub>13</sub>O<sub>2</sub></td><td>6.86</td><td>-</td><td>n.d.</td></tr><tr><td>Perfluorooctanoic acid</td><td>PFOA</td><td>C<sub>8</sub>HF<sub>15</sub>O<sub>2</sub></td><td>7.75</td><td>-</td><td>n.d.</td></tr><tr><td>Perfluorononanoic acid</td><td>PFNA</td><td>C<sub>9</sub>HF<sub>17</sub>O<sub>2</sub></td><td>8.64</td><td>-</td><td>n.d.</td></tr><tr><td>Perfluorodecanoic acid</td><td>PFDA</td><td>C<sub>10</sub>HF<sub>19</sub>O<sub>2</sub></td><td>9.53</td><td>-</td><td>n.d.</td></tr><tr><td>Perfluoroundecanoic acid</td><td>PFUnDA</td><td>C<sub>11</sub>H<sub>21</sub>O<sub>2</sub></td><td>10.42</td><td>-</td><td>n.d.</td></tr><tr><td>Perfluorododecanoic acid</td><td>PFDoDA</td><td>C<sub>12</sub>H<sub>23</sub>O<sub>2</sub></td><td>11.30</td><td>-</td><td>n.d.</td></tr><tr><td>Perfluorotridecanoic acid</td><td>PFTrDA</td><td>C<sub>13</sub>H<sub>25</sub>O<sub>2</sub></td><td>12.20</td><td>-</td><td>n.d.</td></tr><tr><td>Perfluorotetradecanoic acid</td><td>PFTeDA</td><td>C<sub>14</sub>H<sub>27</sub>O<sub>2</sub></td><td>13.08</td><td>-</td><td>n.d.</td></tr><tr><td rowspan="6">PFSAs</td><td>Perfluorobutanesulfonic acid</td><td>PFBS</td><td>C<sub>4</sub>HF<sub>9</sub>O<sub>3</sub>S</td><td>3.68</td><td>-</td><td>n.d.</td></tr><tr><td>Perfluorohexanesulfonic acid</td><td>PFHxS</td><td>C<sub>6</sub>HF<sub>13</sub>O<sub>3</sub>S</td><td>5.25</td><td>-</td><td>n.d.</td></tr><tr><td>Perfluorohexanesulfonic acid</td><td>PFHpS</td><td>C<sub>9</sub>HF<sub>15</sub>O<sub>3</sub>S</td><td>6.14</td><td>-</td><td>n.d.</td></tr><tr><td>Perfluorooctanesulfonic acid</td><td>PFOS<sup>§</sup></td><td>C<sub>8</sub>HF<sub>17</sub>O<sub>3</sub>S</td><td>7.03</td><td>-</td><td>n.d.</td></tr><tr><td>Perfluorodecanesulfonic acid</td><td>PFDS</td><td>C<sub>10</sub>HF<sub>21</sub>O<sub>3</sub>S</td><td>8.81</td><td>-</td><td>n.d.</td></tr><tr><td>Perfluorooctanesulfonamide</td><td>FOSA<sup>§</sup></td><td>C<sub>8</sub>H<sub>2</sub>F<sub>17</sub>NO<sub>2</sub>S</td><td>7.64</td><td>-</td><td>n.d.</td></tr><tr><td rowspan="5">PFOS Prec.</td><td>Perfluorooctanesulfonamidoacetic acid</td><td>FOSAA<sup>§</sup></td><td>C<sub>10</sub>H<sub>4</sub>F<sub>17</sub>NO<sub>4</sub>S</td><td>8.03</td><td>-</td><td>n.d.</td></tr><tr><td>N-Methylperfluorooctanesulfonamidoacetic acid</td><td>MeFOSAA<sup>§</sup></td><td>C<sub>11</sub>H<sub>6</sub>F<sub>17</sub>NO<sub>4</sub>S</td><td>8.81</td><td>-</td><td>n.d.</td></tr><tr><td>N-Ethylperfluorooctanesulfonamidoacetic acid</td><td>EtFOSAA<sup>§</sup></td><td>C<sub>12</sub>H<sub>8</sub>F<sub>17</sub>NO<sub>4</sub>S</td><td>9.34</td><td>-</td><td>n.d.</td></tr><tr><td>N-Methylperfluorooctanesulfonamide</td><td>MeFOSA<sup>§</sup></td><td>C<sub>9</sub>H<sub>4</sub>F<sub>17</sub>NO<sub>2</sub>S</td><td>7.97</td><td>-</td><td>n.d.</td></tr><tr><td>N-Ethylperfluorooctanesulfonamide</td><td>EtFOSA<sup>§</sup></td><td>C<sub>10</sub>H<sub>6</sub>F<sub>17</sub>NO<sub>2</sub>S</td><td>8.50</td><td>-</td><td>n.d.</td></tr><tr><td rowspan="4">FTSAs</td><td>4:2 Fluorotelomersulfonic acid</td><td>4:2 FTSA</td><td>C<sub>6</sub>H<sub>5</sub>F<sub>9</sub>O<sub>3</sub>S</td><td>1.70</td><td>-</td><td>n.d.</td></tr><tr><td>6:2 Fluorotelomersulfonic acid</td><td>6:2 FTSA</td><td>C<sub>8</sub>H<sub>5</sub>F<sub>13</sub>O<sub>3</sub>S</td><td>3.47</td><td>-</td><td>n.d.</td></tr><tr><td>8:2 Fluorotelomersulfonic acid</td><td>8:2 FTSA</td><td>C<sub>10</sub>H<sub>5</sub>F<sub>17</sub>O<sub>3</sub>S</td><td>5.25</td><td>-</td><td>n.d.</td></tr><tr><td>10:2 Fluorotelomersulfonic acid</td><td>10:2 FTSA</td><td>C<sub>12</sub>H<sub>5</sub>F<sub>21</sub>O<sub>3</sub>S</td><td>7.03</td><td>-</td><td>n.d.</td></tr><tr><td rowspan="2">Betaines</td><td>6:2 Fluorotelomersulfonamide akylbetaine</td><td>6:2 FTAB</td><td>C<sub>15</sub>H<sub>21</sub>F<sub>13</sub>N<sub>2</sub>O<sub>5</sub>S</td><td>2.58</td><td>+/-</td><td>n.d.</td></tr><tr><td>8:2 Fluorotelomersulfonamide akylbetaine</td><td>8:2 FTAB</td><td>C<sub>19</sub>H<sub>19</sub>F<sub>21</sub>N<sub>2</sub>O<sub>4</sub>S</td><td>4.03</td><td>+/-</td><td>n.d.</td></tr></tbody></table>

References:
<figure><img src="image_4.png" /><figcaption>**Fig. 2.** Changes in total organic carbon (TOC in %) and relative proportions of three particle size classes in the VRN22-C02 archive and changes in relative density (CT-SCAN images).</figcaption></figure>

have been discontinuous throughout time.

## 3.2. Sediment dating

The onset of nuclear testing, dated to 1955 and marked by the first detection of <sup>137</sup>Cs in the core occurred at a depth of approximately 160 cm, dated to 1954 ± 9 yrs (Fig. 3). The peak of <sup>137</sup>Cs, which is attributed to 1963 in the Northern Hemisphere (peak of nuclear testing fallout), was observed at a depth of approximately 125 cm, and dated with the CF:CS model to 1961 ± 7 yrs. The corresponding <sup>137</sup>Cs mass activity, expressed in Becquerel (Bq) per mass unit, was approximately 65 Bq·kg<sup>-1</sup>, which is higher than the activities generally reported in the literature for sediment cores collected in this region. The presence of <sup>241</sup>Am at depths comprised between 100 and 140 cm (1.5–3 Bq·kg<sup>-1</sup>), produced during atmospheric testing, confirms the origin of the <sup>137</sup>Cs peak. The peak corresponding to the Chernobyl accident in April 1986 was slightly visible in core VRN22-C02 around 29 cm, followed by a sedimentary hiatus between 1991 ± 2 yrs and 2001 ± 2 yrs. This pattern is consistent with the limited Chernobyl fallout in this region as reconstructed by Meusburger et al. (2020).

The measurements of <sup>210</sup>Pb<sub>ex</sub> varied between 38 and 10 Bq·kg<sup>-1</sup> between the top and the bottom of the archive. The decay of this radionuclide was not constant and was separated into two segments, resulting in a change in the sedimentation rate dated approximately to 1975 ± 3 yrs. The age model, which is based on 33 measurements of <sup>210</sup>Pb<sub>ex</sub>, estimated an average rate of sediment accumulation on each of these segments of 1.07 cm·yr<sup>-1</sup> for the recent period (r<sup>2</sup> = 0.97) and 4.19 cm·yr<sup>-1</sup> for the pre-1975 period (r<sup>2</sup> = 0.54).

## 3.3. TrOC occurrences

### 3.3.1. Pharmaceutical products

The vertical distribution of several pharmaceutical product

<figure><img src="image_5.png" /><figcaption>**Fig. 3.** Age model of core VRN22-C02 estimated via a CFCS model. NWT 1963 corresponds to the <sup>137</sup>Cs peak attributed to the atmospheric test maximum, and FF 1955 corresponds to the first detection of <sup>137</sup>Cs. Gray areas mark flood deposits, considered to be instantaneous in the age model.</figcaption></figure>

<footnote>a Chast (2012);</footnote>

<footnote>b Chauvel et al. (2012).</footnote>
<figure><img src="image_6.png" /><figcaption>**Fig. 4.** Sedimentary concentrations of quantified pharmaceutical products in ng·g⁻¹ d.w., white and blurred symbols indicate concentrations < LOQ and <LOD, respectively, gray area marks sedimentary hiatus, yellow stars indicate the years on which the active ingredient in question was supplied on the market, and red stars indicate years of use restrictions enforced.</figcaption></figure>

concentrations is shown in Fig. 4. The highest concentrations were reached for antibiotics such as oxolinic acid and tetracycline, with maximum concentrations of several tens of ng g⁻¹. There is good agreement between the period when the TrOC was first marketed and when it was first quantified in sediments, whereas the time lag between these two periods varied from molecule to molecule. It was particularly important for tetracycline, which was marketed in 1949 and detected for the first time in sediment in 1958.

Regardless of the molecule under consideration and the generation in which it was marketed, similar trends were observed: (i) initial detection/quantification followed by an increase in sediment contamination, followed by (ii) a gradual decrease in contamination. The peak of contamination differed according to the molecules under consideration. For molecules marketed before the 1970s, the contamination peak was generally observed between the late 1970s and the mid-1980s, whereas for more recently marketed molecules, such as oxolinic acid and metoprolol, contamination peaked in the 1980s. Finally, for the most recently marketed drugs (i.e., from the 1990s onwards), such as venlafaxine and tramadol, the contamination peak did not appear to have been reached yet.

Although the peaks in contamination for some molecules appears to be quite old, the corresponding contamination is still quantified in the sediments although at lower levels. This indicates that even though the river water chemical quality has likely improved overall in terms of contamination levels, the diversity of contaminants present is still greater, as already demonstrated in this river basin, further downstream (Thiebault et al., 2021). One antibiotic, sulfamethazine, exhibited a different pattern and was not detected in the sediments deposited over the last 40 years. This likely shows that the use of this molecule was progressively limited and/or abandoned (i.e., limitation for veterinary purpose), as already observed in several archives in the U.S. (Kerrigan et al., 2018; Lara-Martín et al., 2015), which appears to be confirmed in Paris, France, as well. Another exception is trimethoprim, for which the

exclusive use in human medication between 1990 and 2017 was associated with sulfamethoxazole (Thiebault, 2020), and which was not recorded in this archive due to unfavorable properties and ionization state (Table 1). Finally, two molecules displayed anticipated first quantifications in comparison with their marketed year; i) oxazepam, which is the major metabolite of several benzodiazepines that were authorized earlier (Dinis-Oliveira, 2017) and ii) oxolinic acid for which a similar pattern was already identified in a close location in the same river basin (Tamtam et al., 2011). A specificity of pharmaceutical products in comparison with other TrOCs targeted here is that they have been historically sparsely banned or restricted, although their therapeutic uses may vary, either upwards or downwards due to medical practice changes or new perimeters of therapeutic targets. Conversely, the global use of almost all the molecules has increased in recent decades (if we except the use of antibiotics in breeding), which is not reflected in the archive-recorded contamination.

### 3.3.2. Pesticides

Among the 16 targeted pesticides, 9 substances were quantified in the sediment (Fig. 5). The two investigated fungicides (i.e., carbendazim and tebuconazole) were quantified, as were 5 herbicides and two atrazine byproducts. The highest concentrations were found for carbendazim, at concentrations of approximately one hundred ng g⁻¹. As with pharmaceutical products, a good match was found between the year of marketing and the first detection in sediments, with a shorter time lapse between the date of marketing and the first detection. For all the molecules, there was also an increase to a maximum after this first quantification, followed by a decrease in contamination over the most recent years. These peaks were spread over 3 decades, depending on the molecules under consideration: 1980s for atrazine, terbuthylazine and chlortoluron; 1990s for carbendazim and isoproturon; and peak not reached yet for diflufenican and tebuconazole, the two most recent molecules.
<figure><img src="image_7.png" /><figcaption>**Fig. 5.** Concentrations of the quantified pesticides and two atrazine byproducts, hydroxyatrazine (HTZ) and desethylatrazine (DEA), in ng.g⁻¹ in the VRN22-C02 archive, white and blurred symbols indicate concentrations < LOQ and <LOD, respectively, gray area marks sedimentary hiatus. The yellow star indicates the year on which the active ingredient in question was supplied on the market, and the red star indicates the years on which it was banned.</figcaption></figure>

Among the 7 active ingredients considered, 4 were banned between 2003 and 2017, and these bans systematically resulted in a reduction in the sediment concentration of the product under consideration. This decrease also systematically took place before the year of the definitive ban, which can probably be understood as the consequence of partial restrictions on substance use taken before the drastic final decision, as documented for atrazine in 1991 and 1997 (Mahé et al., 2020). Only chlortoluron has shown a marked decline since the 1990s in the absence of a ban. Restrictions were implemented in the 2010s, although there is no evidence of this anticipated decline. This likely reflects possible changes in weed control practices, particularly in favor of molecules such as metolachlor which is currently one of the best-selling products, but was not detected in sediments partly due to limited sediment affinity (Bech et al., 2022).

Atrazine, which is still very frequently measured in surface and deep waters as well as in sediments despite a long-standing ban, was quantified in the current research along with two of its byproducts (deisopropylatrazine was not detected). Another triazine, terbuthylazine was detected with an overlap between market authorization in 1974 and the first quantification of this substance in sediment in 1965. The correspondence of trends for all triazines raises questions about the role of major impurities in the composition of products marketed impacted the patterns for all the substances. The overall trends for the three molecules were identical, with a peak in contamination during the early 1970s and mid-1990s and a parallel decline in levels since then. It is also logical to find that the byproduct hydroxyatrazine (HTZ) is quantitatively the most substantial form, given that after application, it is the form that is dominant in agricultural soils, following the hydrolysis of the parent compound (Jones et al., 1982). The evolution of the ratios between atrazine and its two byproducts (fairly stable for the atrazine/DEA ratio, highly variable for the atrazine/HTZ ratio) could serve as an indicator of residence time and environmental conditions, as it is understood that HTZ is more stable under aerobic conditions than atrazine is and vice versa. The more rapid decline in HTZ compared to atrazine over the recent period therefore likely indicates a release of atrazine from specific environmental stocks, as soils and groundwaters.

### 3.3.3. PFAS

The third group of TrOCs analyzed were PFAS. Unlike the molecules mentioned above, there is little information available on the marketing of PFAS individually, although the first uses were identified in the 1950s and have been steadily increasing since then. The specificity of these molecules is the occurrence of multiple C–F bonds, which are particularly difficult to degrade via natural attenuation mechanisms. All PFAS are considered as meeting the very persistent criteria under the EU REACH regulation since they are either non-degradable or they may transform into persistent transformation products. These different molecules therefore need to be seen as part of a continuum of degradation, particularly of the polar heads of these amphiphilic compounds and the

hydrogenated moiety of polyfluoroalkylated compounds (Buck et al., 2021). Therefore, unlike previous contaminants considered in this research, for which we looked at molecules individually, PFAS can be considered from the perspective of subgroup of molecules (Table S1). The cumulative content of the 42 molecules reached a maximum of 25 ng g⁻¹ that is much higher than the observed French median surface sediments contamination, estimated at a few ng.g⁻¹ (Macorps et al., 2023). Even if this level of sediment contamination is lower than that observed for pesticides or pharmaceutical products, this does not prevent us from observing very contrasting temporal dynamics depending on the type of molecule under consideration (Fig. 6).

To summarize, two main trends were observed: i) PFOS precursors and PFSAs, which have been quantified since 1958 and 1968, respectively, showed an upward trend until 2007, when point contamination decreased overall to much lower levels, and ii) PFCAs and FTSAs, which have been quantified since 1981 and 1995, respectively, for which contamination increased until 2008, both in terms of cumulative concentration and overall contribution to the contaminant signal.

Two key dates likely had an impact on these developments, with the

<figure><img src="image_8.png" /><figcaption>**Fig. 6.** Cumulative concentrations of PFAS sorted by molecule type.</figcaption></figure>
industry first announcing that it would process without PFOS in 2003 (Paul et al., 2009), before its use was considerably reduced by the inclusion of PFOS in the Stockholm Convention on Persistent Organic Pollutants in 2009. The impact of these two measures can be seen in the diversification of PFAS observed from the early 2000s onwards, followed by a sharp decline in PFOS precursors and PFSAs (with maximum PFOS concentrations observed in 2003) from 2003 onwards in sediments (Fig. 8). Through a relay effect, manufacturers turned to other molecules with similar properties and that were considered or perceived to be less problematic from an environmental point of view: FTSAs and Betaines. Their content in sediment was lower than that observed for PFOS precursors. However, 6:2 FTSA was the only molecule showing a constant increase in sediment, indicating that this substitution has potential consequences that are yet unknown. It should be noted that FTSAs may also arise from the transformation of more complex precursors (Field and Seow, 2017). In addition, these compounds can be degraded to PFCAs, the cumulative levels of which have increased over the last decade, particularly for byproducts proven to be produced by aerobic biodegradation of 6:2 FTSA, such as PFHxA, PFHpA and PFPeA (Yan et al., 2024). The trend observed is similar for the two measured betaines (6:2 and 8:2 FTAB).

Although measures to limit the use of PFOS likely have made it possible to limit environmental contamination significantly, the diversity of the several thousand PFAS molecules, some of which, such as TFA, having very limited affinity with sediment, make it difficult to assess the overall environmental quality on the basis of targeted monitoring of large indicator molecules (e.g., 42 in Macorps et al., 2023). However, the trends observed are consistent with data measured in the Rhône river basin in France and Baltic sea (Falandysz et al., 2012; Mourier et al., 2019), and the current investigations clearly reveal the link between past and emerging PFAS contamination and the diversification of quantified compounds (Fig. 7).

### 3.4. Point-source historical impact

As described above, the sedimentary record of past organic contamination makes it possible to retrospectively assess the trajectory of the chemical quality of the Seine River environment. The degree of deterioration in chemical quality is the result of various forcings, particularly those provided by point sources and diffuse pressures. The formers are generated mainly delivered by discharges from wastewater treatment plants and, if we go back in time, by untreated industrial and/or raw domestic wastewater discharges into aquatic environments, which have been banned since the UWD in 1991, but with an application deadline of 10 years.

The history of point source pressures in the Seine River Basin is therefore closely linked to developments in wastewater treatment in general, and Paris wastewater treatment in particular in the case of the VRN22-C02 archive, which was collected downstream of Paris. The wastewater collection system in the Paris conurbation is currently highly concentrated around 6 WWTPs. Of the 530 WWTPs found in the Île-de-France region, these 6 WWTPs collect and treat approximately 80 % of the collected volume. The successive opening of new sections at the Achères site, as well as in the Noisy-le-Grand and Valenton plants, brought the volumes collected and treated closer together (Fig. 8A). The 3 most recently commissioned WWTPs (Colombes, Triel and Blanc-Mesnil, Fig. 8) were set up as part of the deconcentration of treatment and the reduction in capacity at the Achères site and therefore had no major impact on the total collection volumes. Notably, since 2004, almost all the collected wastewater is treated. Treatment technologies, which for a long time focused on carbon pollution, have been adapted since 2007 with the introduction of new treatment processes, particularly biofiltration, for nitrogen purification. Finally, since 2001, clariflocculation treatment has been used to ensure minimal treatment of excess water during intense rainfall. The point-source pressure on the environment in terms of volume discharged without treatment was the highest before 1972 and then gradually decreased until 2004. Since then, any volume collected has theoretically been treated, with the exception of operational accidents or the occurrence of intense rainfall events generating overflows.

Moreover, this drastic improvement in wastewater collection and treatment has an impact not only on conventional treatment parameters but also potentially on TrOCs or at least some of them. For the molecules in question, pharmaceutical product contamination peaked in the 1980s before declining sharply from 2000 onwards. Therefore, despite the very strong increase in use of these substances, the substantial improvement in wastewater collection and treatment, even incomplete (Grandclément et al., 2017), has made it possible to reduce environmental contamination. For molecules that are more likely to be generated by diffuse pollution, such as pesticides, drawing a comparison with the history of wastewater management makes limited sense. For the considered molecules, the decrease in contamination occurred in the 2000s, before being strongly impacted by several bans that led to an improvement in the chemical quality of the sediments. Notably, however, a number of currently widely used molecules, such as prosulfocarb and glyphosate, were not among those monitored here. Finally, in the case of PFASs, which are supplied by multiple sources, the peak in sediment contamination for the list of molecules under consideration was reached in 2003, before declining (Fig. 8B). Therefore, there is a priori a decorrelation between wastewater management and the cumulative content of PFAS,

<figure><img src="image_9.png" /><figcaption>**Fig. 7.** Temporal relative contribution of each PFAS type to total contamination in relative proportion or number of detected molecules (nPFAS), gray area marks sedimentary hiatus.</figcaption></figure>
<figure><img src="image_10.png" /><figcaption>**Fig. 8.** (A) Evolution of collected wastewater volume (in red-brown) in Paris conurbation since WWII and treatment technology, by spreading in purple and by WWTP in green. The evolution of WWTP targets is displayed at the left of the figure. New WWTPs are marked by a yellow star, and (B) sum occurrences of PFAS (blue), pesticides (green) and pharmaceuticals (orange), in ng g⁻¹ d.w. in the archive VRN22-C02, gray area marks sedimentary hiatus.</figcaption></figure>

which has likely been mainly impacted by the 2009 ban on PFOS. In addition, a fluorochemical manufacturing plant in the Oise catchment, a tributary of the Seine River, has been shown on several occasions to have had an impact on the molecular composition of PFAS found in dissolved and particulate phases (Macorps et al., 2022).

### 3.5. Study limitations

Despite the comprehensive nature of this study, several limitations must be acknowledged regarding both the methodological choices and the interpretation of the results. First of all, the respective contributions of sediment and pore water to the observed contamination trends are impossible to disentangle due to the use of freeze-drying for sample preparation. Although this is the standard approach in the field (Ducrocq et al., 2024), separating the two matrices can yield valuable insights (Ahrens et al., 2009), an option that could unfortunately not be implemented in the current research. Sedimentary processes such as adsorption/desorption dynamics, vertical diffusion, and potential post-depositional transformation of mobile TrOCs such as atrazine, although only sparsely observed in specific contexts with very limited sedimentation rates (e.g. Guo et al., 2016), were not explicitly considered in the interpretation of the concentration profiles. The long-term stability of functionalized and moderately polar compounds in sediments remains uncertain, even though the extensive historical data collected from sedimentary archives increases confidence in the overall trends (Heim and Schwarzbauer, 2013).

Discontinuous sedimentation, and particularly the hiatus identified between 1991 and 2001, introduces chronological uncertainty. This

may mask key regulatory or emission-related inflection points, as well as the numerous minor flood events (Fig. 3) that could affect the continuity of organic contaminant deposition. Furthermore, although sedimentological parameters such as total organic carbon (TOC) content and grain size distribution were thoroughly characterized, they were not directly used to normalize contaminant concentrations in the sediment. While TOC values range from 3 % to 15 % along the core, this variability is much smaller in post-1965 sediments (3.2–5.6 %, Fig. 2), which are richer in both diversity and cumulative occurrences of TrOCs (Fig. 8). The added value of organic carbon normalization, tested on two molecules per class (Fig. S1), appears limited, smoothing the contamination pattern for only three layers amongst oldest and sparsely contaminated ones. A key reason for not performing organic carbon normalization lies in the underlying assumption that organic matter governs all sorption and retention processes. This assumption is not universally valid, particularly for ionized molecules (Sigmund et al., 2022). While organic carbon normalization may appear to be more robust, it can distort the bulk signals and lacks a clear justification, which is why it was not applied in this study. All the data required for conducting such a normalization are nevertheless provided if needed in further studies.

### 4. Conclusion

This study assessed the sedimentary occurrence of TrOCs between 1943 and 2021. The sedimentary occurrence of various molecules, i.e., pesticides, PFAS and pharmaceutical products, has been evaluated. These three contaminant classes exhibited contrasted global temporal evolutions that were shown to be mainly impacted by wastewater
management changes (i.e., for point-source contaminant) whereas individual contaminant patterns were strongly impacted by regulatory decisions. The oldest quantification of each molecule was indeed closely related to its market authorization year. A global improvement of sediment chemical quality was observed in the most recent years, as some TrOCs exhibited sharp decrease that can be linked with regulatory decisions, which appeared more or less quickly recorded in sediments depending on the physical-chemical properties of the molecules and their environmental persistence. Overall, this observed improvement in the Seine River chemical quality needs to be moderated, as the targeted list of TrOCs comprised 76 substances belonging to three classes of contaminants representing thousands of molecules, despite the careful selection of environmentally relevant molecules.

A full assessment of the historical chemical quality of the environment is difficult to make, as the sedimentary occurrence of TrOCs does not directly reflect dissolved concentrations. However, the general patterns observed here concur with the various knowledge sources of the historical contamination of molecules (use, point-source management, restriction and ban) at various scales (i.e., local, regional and national). Sedimentary archives can therefore be considered as unique tools in order to retrospectively evaluate the impact of regulatory decision and wastewater management changes on chemical quality, and to reconstruct a posteriori ancient background contamination in case of emerging issue with specific compounds.

### CRediT authorship contribution statement

**Thomas Thiebault:** Writing – review & editing, Writing – original draft, Supervision, Methodology, Investigation, Funding acquisition, Data curation, Conceptualization.
**Pierre Labadie:** Writing – review & editing, Writing – original draft, Methodology, Investigation, Data curation.
**Anthony Foucher:** Writing – review & editing, Writing – original draft, Resources, Methodology, Investigation.
**Jean-Sébastien Barbier:** Writing – review & editing, Methodology, Investigation, Data curation.
**Fabrice Alliot:** Writing – review & editing, Methodology, Investigation, Data curation.
**Thibaut Soubise:** Writing – review & editing, Investigation, Data curation.
**Patrick Pardon:** Writing – review & editing, Methodology, Investigation, Data curation.
**Hélène Budzinski:** Writing – review & editing, Supervision, Methodology.
**Elie Dhivert:** Writing – review & editing, Resources, Methodology, Investigation.
**Olivier Evrard:** Writing – review & editing, Resources, Investigation, Formal analysis.
**Sophie Ayrault:** Writing – review & editing, Supervision, Project administration, Funding acquisition.

### 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

This study was supported by the Sedi-PLAST project funded by the French national agency (ANR-19-CE34-0012) and the PIREN Seine project. F. Baudin and F. Savignac are gratefully acknowledged for their support during Rock-Eval analyses. A. Pavret de la Rochefordière is thanked for its contribution in sediment preparation and analyses. The PLATINE (Organic Environmental Analytical Chemistry Platform) platform at UMR 5805 EPOC and its research engineers are acknowledged for providing the analytical facilities dedicated to PFAS analyses. Finally, we want to acknowledge the three anonymous reviewers for their insightful comments.

### Appendix A. Supplementary data

Supplementary data to this article can be found online at https://doi.org/10.1016/j.jenvman.2025.126276.

### Data availability

Requests to access the data should be directed to the corresponding author.

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