# Ecotoxicological risk assessment using a multi-scale approach in the estuarine copepod Eurytemora affinis applied to an ecdysone agonist

Caroline Arcanjo, Yannick Colin, Nathalie Giusti-Petrucciani, Aurélie Duflot, Jean-Luc Ravanat, Gauthier Trémolet, Romain Coulaud, Anne-Laure Vivant, Benoit Xuereb, Thierry Berthe, et al.

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## To cite this version:

Caroline Arcanjo, Yannick Colin, Nathalie Giusti-Petrucciani, Aurélie Duflot, Jean-Luc Ravanat, et al. Ecotoxicological risk assessment using a multi-scale approach in the estuarine copepod Eurytemora affinis applied to an ecdysone agonist. Ecotoxicology and Environmental Safety, 2025, 304, pp.119081. 10.1016/j.ecoenv.2025.119081. hal-05292809

HAL Id: hal-05292809

https://hal.science/hal-05292809v1

Submitted on 4 Oct 2025

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# Ecotoxicological risk assessment using a multi-scale approach in the estuarine copepod *Eurytemora affinis* applied to an ecdysone agonist

Caroline Arcanjo<sup>a</sup>, Yannick Colin<sup>b,c</sup>, Nathalie Giusti-Petrucciani<sup>a</sup>, Aurélie Duflot<sup>a</sup>, Jean-Luc Ravanat<sup>d</sup>, Gauthier Trémolet<sup>a</sup>, Romain Coulaud<sup>a,d</sup>, Anne-Laure Vivant<sup>b</sup>, Benoit Xuereb<sup>a,d</sup>, Thierry Berthe<sup>b,c</sup>, Joëlle Forget-Leray<sup>a,d</sup>, Céline Boulangé-Lecomte<sup>a,* e</sup>

## ARTICLE INFO

Keywords:

Behavior

Copepod

Epigenetics

Estuaries

Larval development

Microbiota

## ABSTRACT

Previous transcriptomic analyses have demonstrated that exposure to tebufenozide (TEB) induces significant alterations in gene expression in the crustacean *Eurytemora affinis* (calanoid copepod; *E. affinis*), supporting its suitability as a model organism for investigating the ecotoxicological effects of ecdysone agonists in estuarine environments. Based on these findings, we hypothesized that TEB exposure may disrupt biological processes across multiple levels of organization in *E. affinis*. We conducted an integrated assessment of upstream (epigenetics) and downstream (locomotion, larval development) endpoints through multiple assays. Additionally, we examine the indirect effects of TEB by evaluating the changes in the copepod microbiota. We showed that TEB had no impact on global DNA methylation (HPLC-MS/MS) in both males and females. In contrast, significant effects were observed on the swimming behaviour of males, with copepods exposed to TEB being more affected by the light/dark transition and showing hyperactivity after 4 days. The larval development was also shown to be accelerated by TEB after 4 days. Finally, significant changes in microbiota were highlighted, including a decrease in both alpha and beta diversity indices, along with an increase in the abundance of the Gammaproteobacteria class. The effects observed in *E. affinis* show that TEB could impact fitness and increase vulnerability of non-target organisms. The strategy of prioritising the responses to be investigated on the basis of gene expression has proven effective. The study reinforces the relevance of using *E. affinis* for environmental risk assessment in estuaries, as it can be used to assess multiple endpoints at different biological scales.

## 1. Introduction

Copepods (Crustacea: Copepoda) are zooplanktonic organisms widely widespread from freshwater to marine environments playing a pivotal role in the food web being the link between primary producers and higher predators. As such, copepod species are relevant model organisms for ecotoxicological studies (Kulkarni et al., 2013a, Kwok et al., 2015). Among them, the calanoid copepod *Eurytemora affinis* has been used to assess the impact of environmental pollution in laboratory experiments. As well as displaying classic model organism characteristics (e.g. sexual dimorphism, short generation time, ubiquity or ecological role) (Kwok et al., 2015), *E. affinis* is a relevant species for transition water, for which few model species are available. Its sensitivity towards

environmental pollution has been studied from single-substance to environmental matrix exposures across endpoints including reproduction (Forget-Leray et al., 2005, Das et al., 2022), development behaviour (Cailleaud et al., 2011a, Arcanjo et al., 2023) or gene expression (Legrand et al., 2016, Lee et al., 2018, Arcanjo et al., 2021).

Substances of concern include insect growth regulators (IGRs), which target insect growth and development, and for which risk assessment is of interest in crustaceans. IGRs are described as selective insecticides inhibiting or interfering with hormones (i.e. juvenile and moulting hormone) or proteins (i.e. chitin) involved in insect metamorphosis and playing a role in reproduction (Tunaz and Uygun, 2004). Despite the high selectivity of ecdysteroid and juvenile hormone insecticides toward their target insect species (Smagghe et al., 1998),

<footnote><sup>a</sup> Université Le Havre Normandie, Université de Reims Champagne-Ardenne, INERIS, Normandie Univ, UMR-I 02 SEBIO, Le Havre F-76600, France</footnote>

<footnote><sup>b</sup> Univ Rouen Normandie, UNICAEN, CNRS, M2C UMR 6143, Rouen F-76000, France</footnote>

<footnote><sup>c</sup> Sorbonne Université, CNRS, EPHE, UMR 7619 METIS, Paris F-75005, France</footnote>

<footnote><sup>d</sup> University of Grenoble Alpes, CEA, CNRS, IRIG, SyMMES UMR 5819, France</footnote>

<footnote>* Corresponding author.</footnote>

<footnote>E-mail address: celine.lecomte@univ-lehavre.fr (C. Boulangé-Lecomte).</footnote>
concerns persist regarding their potential impact on non-insect arthropods, such as crustaceans, which share conserved endocrine pathways involved in moulting. Thus, in a previous study, we assessed the effects of tebufenozide (TEB) as an IGR model on *E. affinis* gene expression (*Arcanjo et al., 2021*). This insecticide is an ecdysone receptor (EcR) agonist acting as a direct ligand to EcR mimicking 20-hydroxyecdysone (20E) moulting hormone and leading to a lethal precocious moult in its targeted species, i.e. lepidopteran (*Chandler et al., 1992*, *Song et al., 2017*). TEB is described as presenting a very low hazard for non-target species (*Carlson, 2000*). In crustaceans, its binding capacity to EcR was assessed *in vitro* and revealed a weak to moderate EcR activation (*Kato et al., 2007*, *Yokota et al., 2011*, *Tarrant et al., 2011*). In *E. affinis*, our previous study showed that early responsive genes (i.e. Broad complex) triggered by 20E or 20E-agonist compounds were unexpectedly down-regulated suggesting the loss of the EcR activation (*Arcanjo et al., 2021*). Nevertheless, gene expression modulations were observed in adult copepods after 3 days of exposure, with some being directly involved in the moulting pathway (i.e. early and late responsive genes) and others being identified as a general stress response (i.e. methylation, defence response). This previous study also highlighted some putative TEB effects at higher biological levels that could be interesting to study, i.e. locomotor behaviour linking to modulation of genes involved in muscle contraction or neurotransmission, DNA methylation and moulting and metamorphosis impairments.

Based on these findings, we hypothesized that exposure to TEB, used as a model IGR insecticide, may disrupt biological processes across multiple levels of organization in *E. affinis*. The endpoints were selected on the basis of the effects highlighted at the gene level in our previous study (*Arcanjo et al., 2021*). The swimming behaviour of adult copepods was investigated under dark/light phases known to trigger locomotor activity in crustaceans (*Kohler et al., 2021, 2018; Love et al., 2020*). Adult copepods were also used to address the effects on global DNA methylation, an epigenetic marker generally associated with gene expression silencing (*Vandegehuchte and Janssen, 2011*). In invertebrates, DNA methylation was shown to be modulated by anthropic pollution (reviewed in Srut 2021) thus being a relevant tool to assess their effects. The impact on the moulting cycle was assessed by exposing copepod larvae (i.e. nauplius stage) to follow their development over time. Finally, in a context where the integration of microbial ecotoxicology into environmental risk assessment is gaining increasing attention, we investigated the potential disruption of the copepod-associated microbiota. In invertebrates, the microbiota is known to regulate key biological functions such as growth, digestion, immunity, stress tolerance, and detoxification (*Holt et al., 2021*, *Akbar et al., 2022*, *Hella et al., 2023*). In copepods, gut-associated microorganisms have been suggested to facilitate nutrient assimilation by degrading complex organic matter such as phytoplankton and detritus (*De Corte et al., 2018; Chae et al., 2021; Hirano et al., 2024*). Moreover, specific microbial taxa may enhance host resilience under environmental stress, for example by degrading cyanobacterial toxins during harmful algal blooms (*Gorokhova et al., 2021*). Bacteria colonizing the exoskeleton can also contribute to host defence by producing antimicrobial compounds that inhibit pathogens or fouling organisms (*Nagasawa et al., 1987*). These functions underline the central role of the copepod microbiota in host health and ecological performance, and suggest that its alteration may have significant consequences in contaminated environments.

## 2. Materials and methods

### 2.1. Copepod sampling and stabling

Copepods were sampled at ebb tide in the oligo-mesohaline zone of the Seine estuary at two sampling stations 20 km apart (Villequier 49°30'51.6"N 0°40'36.9"E or Tancarville 49°28'30.5"N 0°27'48.5"E, Normandy, France) (*Mouny and Dauvin, 2002*). Copepods were

collected using a plankton net (200 µm), washed through successive sieves to eliminate particles and predators, and finally collected on 500 µm and 250 µm mesh size sieves. To ensure a sufficient density of copepods for each endpoint measured (see the following materials and methods sections), one sampling per endpoint was performed. Samplings were performed in July 2019 (microbiota), September 2020 (DNA methylation), June 2021 (swimming behaviour) and January 2023 (growth). Studying wild populations from different samples enables to consider the genetic background of individuals, thereby avoiding the overestimation of responses, as frequently observed with laboratory populations (*Romero-Blanco and Alonso, 2022*). Copepods were transported to the laboratory in cool boxes containing Seine water and placed in a thermoregulated room at 15°C where salinity was adjusted dropwise to 15 psu (practical salinity unit) with a mixture of UV-treated filtered seawater (Deauville, Normandy, France) and dechlorinated tap water. When temperature and salinity were adjusted, copepods were transferred to a clean 15 psu tank using 355 µm mesh size sieves for adult-only collection. Throughout the husbandry, which lasts three days before exposure, copepods were fed *ad libitum* with *Rhodomonas salina* algae and kept under optimal conditions (i.e., 15 °C, 15 psu, 12 h:12 h light-dark cycle, *Devreker et al., 2009*). Standardised acclimatisation conditions help to minimise potential variability between samples related to the physiological state of individuals, and enable the deposition of contaminants in *E. affinis* (*Cailleaud et al., 2011b*).

### 2.2. Chemical preparation and exposures

Tebufenozide (TEB, CAS Number 112410–23–8, analytical standard, Sigma-Aldrich) was diluted in acetone (Sigma-Aldrich) to get stock solutions matching the concentrations needed for each endpoint assessed in the present study (see the following materials and methods sections). All experiments were performed with a final acetone concentration of 0.01 % (according to OECD, 2019) in 15 psu filtered (0.45 µm) and oxygenated water. Consequently, a control solvent (0.01 % acetone i.e. the final acetone concentration in TEB-doped conditions) was performed for each experiment. A saturation period of experimental units with the exposure solutions of at least three days before the beginning of experiments was respected for each endpoint. The experimental units were pyrolysed glass beakers, except for the larval bioassay, where sterile 24-well plates (Nunc™ Non-Treated Multidishes, Thermo Scientific) were used. Maximum copepod density was 1 copepod per millilitre in each experimental unit. Exposure solutions were daily renewed (50 % of the media). The following sections will describe the specificity of exposure methods for each endpoint assessed in the present study. The experimental design is described for each endpoint in Fig. 1. Water samples were collected for all endpoints, except for the larval bioassay due to the small exposure volumes, and analysed by HPLC-MS (La Drôme Laboratoire, Valence, France).

### 2.3. DNA methylation

Adult *E. affinis* were exposed to 50 µg/L of TEB in triplicate. This concentration has been shown to significantly affect gene expression in *E. affinis* after a 3-day exposure (*Arcanjo et al., 2021*). This corresponds to the maximum expected concentration in water (lake) after TEB application in a nearby forest (*Kreutzweiser and Thomas, 1995*). Concentrations measured in the environment (surface water) vary, ranging from nanograms to micrograms per liter (*Comoretto et al., 2007; Ngin et al., 2024*). In particular, surface water near orchards regularly reports concentrations of 20 µg/L (*Süß et al., 2006*). Copepods were fed daily *ad libitum* with *Rhodomonas salina* (15,000 cells/mL) one hour before the water renewal. After 2, 3 and 8 days of exposure, pools of 30 males or 20 non-ovigerous females per triplicate were made to obtain the required quantity of DNA for the HPLC-MS/MS approach (1 µg). Briefly, males and non-ovigerous females were sorted under a stereomicroscope and placed in 1.5 mL collection tubes. Exposure water was gently removed
<figure><img src="image_4.png" /><figcaption>**Fig. 1.** Experimental design representing each endpoint assessed, with sampling time in days, tebufenozide (TEB) concentrations and copepod stage. The ♀ symbol represents non-ovigerous females.</figcaption></figure>

and pools were washed with nuclease-free water. Nuclease-free water was gently removed and collection tubes were quickly freeze-dried in liquid nitrogen before being stored at -80°C until used. DNA extraction was performed with QIAamp DNA micro kit (Qiagen) following manufacturer instructions. The global DNA deoxycytidine methylation (% of 5mdC relative to dC) quantification by HPLC-MS/MS as described by Cribiu and colleagues (Cribiu et al., 2018).

## 2.4. Swimming behaviour

The swimming behaviour was assessed on males because they are known to be particularly sensitive to stress (Li et al., 2019, 2020, Kadiene et al., 2020, Colin et al., 2023) and can be unambiguously distinguished from females and copepodites by their geniculated antennae. Adult male *E. affinis*, sorted under a stereomicroscope, were exposed to 50 µg/L of TEB. They were fed daily *ad libitum* with *R. salina* (15,000 cells/mL) one hour before the experimental unit water renewal. Swimming behaviour (i.e. velocity) was assessed as previously described (Arcanjo et al., 2023). Briefly, swimming behaviour was assessed by video tracking using the DanioVision observation chamber connected to the EthoVision XT 15 software (Noldus, Netherlands). After 3 and 4 days of exposure, eight unspoiled and active males were placed in the central position of a 24-well plate (4 control and 4 exposed) containing 1 mL of exposure water. Eight plates were successively analysed to track 32 individuals per condition. Plates were prepared extemporaneously to the beginning of the video tracking. After a 6 min of acclimatization in the dark in the observation chamber, copepod movements were recorded during 2 cycles of 2 min dark/2 min light. Raw data were used to extract velocity (mm/s). From velocity data, the 3 swimming states described by Michalec and colleagues were distinguished namely (i) the break state (velocity < 1 mm/s) for copepods that do not swim or whose activity is minimal to avoid sinking, (ii) the cruising state (1 ≤ velocity ≤ 5 mm/s) when copepods swim actively and (iii) the fast state (velocity > 5 mm/s) for copepods that swim with swift movements (Michalec et al., 2013). The length of each copepod, which could represent a confounding factor, was determined using the ImageJ software (version 1.53q) from a photo taken at the end of the recording (Leica MZ10F, LAS V4.13).

## 2.5. Larval bioassay

The larval bioassay procedure was adapted from Lesueur and colleagues (Lesueur et al., 2013, 2015). Two days before larvae exposure, *E. affinis* ovigerous females were sorted under a stereomicroscope and placed in egg-laying devices (mesh size 200 µm) in 1.5 L of 15 psu filtered (0.45 µm) and autoclaved water under constant oxygenation. They were fed daily with *R. salina* (15,000 cells/mL) until the beginning of the larvae bioassay. After two days, nauplii were collected and sorted under a stereomicroscope. A pool of 50 nauplii was stored in 95 % ethanol/5 % glycerol at 4°C to determine the mean initial length as described below. Others were placed in well plates containing medium i. e. control (water), control solvent or TEB (5 ng/L, 50 ng/L, 500 ng/L) at a rate of 50 nauplii per condition. The TEB concentrations were chosen

as dilutions (0.0001X, 0.001X and 0.01X) of the maximum concentration tested on adults in order to avoid mortality, since the larval stages of copepods are reportedly more sensitive than the adult stages (Kulkarni et al., 2013b). These concentrations correspond to the lower end of the range of concentrations found in the environment, which extends from nanograms to micrograms per liter (Comoretto et al., 2007; Ngin et al., 2024). The Nauplii were fed with *R. salina* (15,000 cells/mL) at the beginning of the exposure. After 2 days, dead nauplii were counted and removed from the wells. The media were renewed with extemporaneously prepared solutions and nauplii were fed with *R. salina* (15,000 cells/mL). Four days after the beginning of nauplii exposure, dead nauplii were counted, and living nauplii were pooled by condition to be stored in a 95 % ethanol / 5 % glycerol mixture at 4°C until used for length assessment. Each nauplius was measured using the ImageJ software (version 1.53q) from photos (Leica MZ10F, LAS V4.13). The larval length at the end of the experiment was normalized by the mean initial length. Nauplii developmental stages were deduced from the nauplius length as described by Katona (Katona, 1971).

## 2.6. Microbiota

For microbiota analysis, total DNA was extracted from pools of 25 adult copepods using the DNeasy® Blood and tissue kit (Qiagen), following manufacturer's instructions with some modifications as described elsewhere (Colin et al., 2023). The sample pools consisted of copepods of the same sex (males or non-ovigerous females), either untreated (control solvent) or treated with TEB 0.5 µg/L or TEB 50 µg/L for 3 days (Table S1). These concentrations have been shown to differentially affect gene expression in *E. affinis* (slight vs strong; Arcanjo et al., 2021). DNA concentration and quality were assessed with a Nanodrop spectrophotometer (NanoDrop™ 2000/2000c, Thermo Scientific). High-throughput 16S rRNA gene paired-end sequencing was performed at MrDNA Molecular Research LP (Shallowater; TX, USA) on the Illumina MiSeq platform. The 16S rRNA libraries were prepared using the PCR primers 341 F/785 R amplifying the V3-V4 hypervariable region (Klindworth et al., 2013) with standard Illumina barcodes and adapters. Raw sequences were quality-filtered using the dada2 R package (v.1.29.0) (Callahan et al., 2016). Amplicon sequence variants (ASVs) were defined and taxonomies were assigned using the SILVA rRNA reference database (v.138). After quality-filtering, denoising and merging of paired-end reads, a total of 458,658 high-quality non-chimeric reads were retained (Table S1). Sequences were rarefied to an even sequencing depth (13,104 reads per sample) to account for unequal sequencing depth across samples. Both alpha-diversity and beta-diversity metrics were calculated in R using the phyloseq package (v.1.32.0) (McMurdie and Holmes, 2013). The raw 16S rRNA gene sequences were deposited in the European Nucleotide Archive and are accessible under project accession number PRJEB82620.

## 2.7. Statistical analysis

Statistical analyses were performed with the R software (v4.2.1, R
Core Team 2022). The length of the males whose swimming behaviour was studied was compared between the control and TEB groups for each experiment and time point. After checking for test assumptions, the mean lengths were compared by Student or Wilcox tests. No differences were highlighted; thus, copepod length was not included in the model for velocity analysis. The velocity data were analysed in a 2-min time bin using a linear mixed-effect model with treatment (control, TEB) and phase (dark, light) as fixed effects and subject identifiers as random effects to correct for repeated measures between dark and light phases. Model residuals were checked for ANOVA test assumption followed by a pairwise analysis. The swimming state data were analysed using repeated measures ANOVA for within-condition comparison to consider repeated measures on the same subject. To test the influence of TEB on larval stages, a multinomial logistic regression model (multinom function in the package *nnet*) was fitted and a deviance analysis procedure (R package *car*) was used. Data from other assessed endpoints were analysed by paired (Student or Wilcoxon) or multiple (ANOVA or Kruskal-Wallis) comparisons of mean or proportion, after checking for test assumptions. Multiple comparisons were followed by pairwise comparison tests (Student or Tukey). For each endpoint, comparisons with p-values ≤ 0.05 were considered significant.

# 3. Results

## 3.1. Tebufenozide concentration in water

Water samples were collected at the beginning of each experiment (day 0) and at each time point for the swimming behaviour (days 2 and 4), DNA methylation (days 2, 3 and 8), and microbiota (day 3) (Table 1). In all experiments, no TEB was detected in water samples from control groups. The concentrations in the samples from the lowest nominal TEB concentration (0.5 µg/L; microbiota experiment) were under the quantification limit for both time points (day 0 and day 3). For the highest nominal TEB concentration (50 µg/L), measured concentrations were constant along the DNA methylation (day 0 = 30.63, day 2 = 30.23, day 3 = 28.98 and day 8 = 34.11 µg/L) and the microbiota (day 0 = 13.82 µg/L and day 3 = 11.73 µg/L) experiments. Regarding the swimming behaviour, measured concentrations were constant from day 0 to day 2 and reached a value close to the nominal expected concentration of 50 µg/L at day 4 (50.45 µg/L).

## 3.2. Global deoxycytidine methylation

The exposure duration (2, 3 or 8 days) to TEB did not affect the global deoxycytidine methylation level compared to the control for both non-ovigerous females and males (Student tests, p > 0.05, Fig. 2A). As a result, control and TEB data were pooled to express the basal methylation level for both sexes. The basal methylation level was higher in non-ovigerous females compared to males over the time spent in the laboratory (Student tests, p < 0.05) except for the 8-day time point (Wilcoxon test, p > 0.05, Fig. 2B). While the basal methylation level in non-ovigerous females did not vary along the time spent in the laboratory

(Kruskal test, p > 0.05), males showed a basal methylation level significantly lower after 2 days compared to 3 days and 8 days in the laboratory (ANOVA, pairwise comparison Student test, p < 0.05, Fig. 2B). Furthermore, the data dispersion, calculated as the interquartile range (IQR), appeared to be more important in non-ovigerous females compared to males at day 2 (non-ovigerous females IQR = 6 x males IQR). This indicator tended to equalise between non-ovigerous females and males at day 3 (non-ovigerous females IQR = 1.5 x males IQR) and day 8 (non-ovigerous females IQR = 1.1 x males IQR).

## 3.3. Velocity and swimming state

The male copepod swimming velocity was assessed after 3 and 4 days of exposure. The analysis revealed that for both exposure periods, 68 % of the variance was explained by the subject variable. A significant effect of TEB was observed after 4 days, with exposed copepods swimming faster than controls (LME, p < 0.01, Fig. 3A). In both exposure periods, a significant effect of the dark/light phases was highlighted, with copepods swimming faster during dark phases regardless of exposure condition (LME, p < 0.001, Fig. 3A). At day 3, the analysis revealed a significant effect of the treatment and dark/light phase interaction, with TEB-exposed copepods being more affected by the dark/light phase transitions (LME, p < 0.01, Fig. 3A). Break and cruising were highlighted to be the main states involved in the swimming activity (rMANOVA, p < 0.01, Fig. 3B). When comparing control and TEB data for each state across conditions, no difference was observed at day 3 (Student test, p > 0.05). In contrast, after 4 days, copepods in the TEB condition were characterized by less break states (Student test, p < 0.05) compared to the control, in favour of cruising and fast states (Student test, p < 0.001).

## 3.4. Larval survival, growth and stages

Nauplii survival was assessed after 2 and 4 days of exposure to 5, 50 and 500 ng/L. At the end of the experiment (day 4), survival in solvent controls was not significantly different from the recommended threshold for acute and chronic tests (i.e. 80 %; Ziegenfuss and Hall, 1998; data not shown). No significant difference in survival was observed between the control solvent and the TEB concentrations for each time point (proportion test, p > 0.05; data not shown). Neither the solvent (i.e. acetone, used to dissolve TEB) nor the three concentrations of TEB had any significant effect on the larval length, compared to the control (ANOVA, pairwise comparison Tukey, p > 0.05, Fig. 4A). However, significant differences in the distribution of the nauplii stages were observed between conditions (deviance analysis, p < 0.01). After 4 days, the larval development was significantly faster in organisms exposed to TEB than in the control groups, with no larvae in stage 2 and a high proportion of larvae in stages 5 or 6 (56.6 %, 42.0 % and 46.2 % for 5, 50 and 500 ng/L of exposure respectively; Fig. 4B). In contrast, in the controls the large majority of larvae were in stages 2, 3 or 4 (86.9 % and 82.6 % for Control and Control solvent respectively; Fig. 4B) after 4 days.

<table><caption>**Table 1**
concentration of tebufenozide in water samples for each endpoint and exposure duration.</caption><thead><tr><th rowspan="2">Endpoint</th><th rowspan="2">Condition</th><th rowspan="2">Nominal concentration (µg/L)</th><th colspan="5">Measured concentration (µg/L)</th></tr><tr><th>Day 0</th><th>Day 2</th><th>Day 3</th><th>Day 4</th><th>Day 8</th></tr></thead><tbody><tr><td rowspan="2">Behaviour</td><td>Control</td><td>0</td><td>&lt; QL</td><td>&lt; QL</td><td>NA</td><td>&lt; QL</td><td>NA</td></tr><tr><td>TEB</td><td>50</td><td>23.17</td><td>17.86</td><td>NA</td><td>50.45</td><td>NA</td></tr><tr><td rowspan="2">Methylation</td><td>Control</td><td>0</td><td>&lt; QL</td><td>&lt; QL</td><td>&lt; QL</td><td>NA</td><td>&lt; QL</td></tr><tr><td>TEB</td><td>50</td><td>30.63</td><td>30.23</td><td>28.98</td><td>NA</td><td>34.11</td></tr><tr><td rowspan="3">Microbiota</td><td>Control</td><td>0</td><td>&lt; QL</td><td>NA</td><td>&lt; QL</td><td>NA</td><td>NA</td></tr><tr><td rowspan="2">TEB</td><td>0.5</td><td>&lt; QL</td><td>NA</td><td>&lt; QL</td><td>NA</td><td>NA</td></tr><tr><td>50</td><td>13.82</td><td>NA</td><td>11.73</td><td>NA</td><td>NA</td></tr></tbody></table>

NA: not assessed; QL: quantification limit (0.4 µg/L)
<figure><img src="image_5.png" /><figcaption>**Fig. 2.** Global deoxycytidine methylation level (%) in non-ovigerous females and males after exposure to 50 µg/L of tebufenozide for 2, 3 and 8 days. DNA was extracted from pools of 30 males (n = 3 per condition) or 20 non-ovigerous females (n = 3 per condition). The analyses were conducted before (A) and after (B) control solvent and tebufenozide-exposed groups were pooled. Error bars represent standard error.</figcaption></figure>

### 3.5. Microbiota

Exposure of copepods to the lowest TEB concentration (0.5 µg/L) did not lead to any significant shift in the microbiota, whereas higher concentration (50 µg/L) resulted in marked changes in both alpha and beta diversities. The microbial richness and diversity metrics were reported to decrease with increasing TEB concentrations (Fig. 5A-C). Based on a principal coordinate analysis carried out at ASV level, the first axis (89 %.1 of total variance) showed a clear dichotomy between the microbial community profiles exposed or not to 50 µg/L TEB (Fig. 5D). The effect of TEB on the bacterial community structure was further supported by a permutational multivariate analysis of variance (PERMANOVA, F = 62 p < 0.001). Taxonomic classification of ASVs further revealed that changes in the microbial community structure were mainly attributed to an increase in the relative abundance of the *Gammaproteobacteria* members at the expense of the *Alphaproteobacteria* in copepods exposed to 50 µg/L TEB (Fig. 5E). No significant differences were observed between the microbiota of female and male individuals.

## 4. Discussion

Based on a previous evaluation of the TEB effects on *E. affinis* gene expression, this study evaluates the impact of this ecdysone agonist both upstream *i.e.* on the global DNA methylation and downstream *i.e.* on the individual level by investigating adult behaviour and larval development. We extended our analysis beyond the copepod by studying its microbiota.

### 4.1. Epigenetics

Epigenetic modifications are reversible changes in gene expression that allow organisms to adapt to their environment in response to external factors such as changes in temperature or pollution. This mechanism plays a crucial role in influencing the phenotype of individuals in the short or long term. Monitoring these changes represent a promising tool for tracking the effects of environmental stress. Among the epigenetic analytical approaches, DNA methylation appears to be the most robust epigenetic regulatory mark studied in aquatic animals (Pham et al., 2023). In the present study, the global DNA methylation was not significantly affected in response to TEB exposure, regardless of copepod sex (male and non-ovigerous female). An average low percentage of DNA methylation, around 1 %, was observed in both sexes,

which is consistent with the average DNA methylation levels reported in invertebrates (Cribiu et al., 2018, Reviewed in Srút 2021). Studies reported hypomethylation after chemical exposure, i.e. endocrine disruptor vinclozolin and cadmium, in crustaceans (Vandegehuchte et al., 2010, Cribiu et al., 2018). One hypothesis to explain this hypomethylation is a greater involvement of homocysteine in the glutathione synthesis for xenobiotics biotransformation rather than the DNA-methyl-donor-via-DNMT SAM regeneration through the methionine cycle (Lee et al., 2009). Interestingly, our previous transcriptomic analysis on *E. affinis* after TEB exposure at the same concentrations reveals a down-regulation of genes coding for DNMT activation and glutathione hydrolase, while genes coding for glutathione S-transferases were up-regulated in males (Arcanjo et al., 2021). This finding could explain the hypomethylation pattern observed in males compared to females in the present study according to the hypothesis by Lee and colleagues (Lee et al., 2009). This global DNA methylation pattern in males returns to that in females after the longest period in the laboratory. Thus, we could hypothesise that the pattern of methylation observed in the present study could be related to the sex-specific differential sensibility to stress already reported in copepods, i.e. males being more sensitive than females (Li et al., 2019, 2020, Kadiene et al., 2020, Colin et al., 2023), more than a sex-specific DNA methylation pattern *per se*. The sexual dimorphism observed in the DNA methylation profiles of *E. affinis* does not appear to be generalised among crustaceans, since no differences have been highlighted by Cribiu and colleagues in male and female gammarids using the same analysis technique (Cribiu et al., 2018). Nevertheless, it is interesting to continue investigating DNA methylation patterns according to sex, as differential DNA methylation could be found at gene-specific regions or tissues (reviewed in Šrut., 2021). To our knowledge, few studies have investigated epigenetic marks as DNA methylation on copepod species under environmental stress. The effects of ocean acidification, nanoplastic, and chemical stress on DNA methylation patterns in copepods have been investigated under single, transgenerational, and multigenerational exposure scenarios (Guyon et al., 2018, Aluru et al., 2021, Lee et al., 2022, 2023). Consistent with the findings of the present study, these previous studies reported no significant impact of such exposures on global DNA methylation levels (Guyon et al., 2018, Aluru et al., 2021, Lee et al., 2023). This does not preclude the possibility of subtle changes occurring in specific organs or genome regions, which would be obscured by global DNA methylation assessments in whole organisms, even more so for small organisms that must be pooled. In accordance,
<figure><img src="image_6.png" /><figcaption>**Fig. 3.** Swimming behaviour of males (n = 32 per condition) after exposure to 50 µg/L of tebufenozide for 3 and 4 days, focusing on velocity (A) and state duration (B). Error bars represent standard error.</figcaption></figure>

differential DNA methylation patterns were observed on specific
genome regions emphasising the role of DNA methylation in environ-
mental stress response in copepods (Aluru et al., 2021, Lee et al., 2022,
2023). Those differentially methylated regions were also associated with
specific genes involved in biological function impaired by exposures (i.e.
reproductive impairments) suggesting the link between epigenetic
modifications and phenotypic impairments in copepods (Lee et al.,
2022, 2023). Together, these studies position DNA methylation as a
promising biomarker for assessing ecotoxicity in copepods, as well as in
other invertebrates (Srut., 2021), while also emphasizing the importance
of further research, especially through targeted approaches.

4.2. Locomotion

The alteration in the expression of genes involved in biological

functions such as moulting, muscle contraction, visual perception or
neurotransmission, highlighted in our previous study, drew our atten-
tion to the possible effect of TEB on copepod locomotor behaviour
(Arcanjo et al., 2021). Thus, in the present study, we assessed the ve-
locity and swimming state of male *E. affinis*. Regardless of exposure time
and treatment, male *E. affinis* swimming velocity was higher in the dark
phases compared to the light ones, in agreement with previous studies
on artemia and *E. affinis* (Kohler et al., 2021, Arcanjo et al., 2023). The
hypothesis behind this observation could be related to the foraging
behaviour of copepods that performed diel vertical migration to avoid
predation during daytime (Leising et al., 2005). A significant interaction
between treatment and dark/light phases was observed on day 3,
highlighting that TEB-exposed copepods were more affected by the
phase transition than controls. Moreover, the exposure to TEB induced
significant hyperactivity after 4 days, responsible for maintaining the
<figure><img src="image_7.png" /><figcaption>**Fig. 4.** Nauplii normalized length (A) and stage proportion (B) after exposure to tebufenozide at 5, 50 and 500 ng/L for 4 days. The number of individuals measured ranged from 21 to 31. Error bars represent standard error.</figcaption></figure>

<figure><img src="image_8.png" /><figcaption>**Fig. 5.** Impact of TEB exposure on the microbiota diversity of the copepod Eurytemora affinis. (A–C), alpha diversity indices (richness, Shannon, and Simpson) between control and exposed groups (0.5 and 50 µg/L). (D), Principal Coordinates Analysis (PCoA) based on Bray-Curtis dissimilarity, highlighting shifts in community composition at the ASV level. (E), relative abundance of major phyla, with Pseudomonadota further resolved at the class level. Point shapes distinguish male (diamonds) and female (circles) individual pools. Pools (n = 3 per condition) were prepared from 25 individuals.</figcaption></figure>

same level of speed compared to controls in which it tends to reduce. Regarding the swimming state, it seems that break and cruising were the main states observed, regardless of the condition, in agreement with other studies on *E. affinis* (Cailleaud et al., 2011a; Michalec et al., 2010, 2013). A reduction in the duration of the break phase after 4 days of TEB exposure compared with controls explain the hyperactivity observed.

Hyperactivity after exposure to chemicals was already observed in crustaceans including *E. affinis* (Michalec et al., 2013; Soose et al., 2023). It could be responsible for over-consumption of energy reserves and an increased inability to reproduce, feed and avoid predation, which could affect the survival and dynamics of populations (Amiard-Triquet et al., 2013). Further investigations, including females and life stages,
are needed to understand the effects of chemicals on the copepod swimming behaviour. Nonetheless, the development of novel approaches in recent years to help the standardization of behavioural ecotoxicology assessment in invertebrates (Kohler et al., 2021, Soose et al., 2023) encourages studies in this area and will promote the use of such endpoint in chemical risk assessment (Ågerstrand et al., 2020, Ford et al., 2021).

### 4.3. Larval development

In the present study, a larval bioassay at TEB sub-lethal concentrations was performed to assess its effect on *E. affinis* nauplii growth and development. No solvent (i.e. acetone) or TEB effects on nauplii length were observed after 4 days. However, the TEB exposure resulted in the presence of a majority of late stages (i.e. N5 and N6) and the absence of N2 stages, unlike the controls, suggesting an acceleration in the development of the nauplii. This acceleration could lead to an increased metabolic cost, which could impact individuals' fitness. The developmental acceleration could be a phenotypic response to the previously reported effects on the expression of genes encoding proteins involved in moulting in *E. affinis* (Arcanjo et al., 2021). These results are consistent with the mode of action of TEB, which acts as an agonist of ecdysteroids and accelerates moulting in insects (Chandler et al., 1992, Retnakaran et al., 2001). Interestingly, TEB exposure was reported to have no effect on moulting in the juvenile *Neomysis integer* shrimp exposed to high concentration i.e. 100 µg/L, which could highlight the importance of the exposure window and/or the interspecies variability (De Wilde et al., 2013). Thus, the *E. affinis* larvae appear to be particularly sensitive, confirming the relevance of using this species for ecotoxicological risk assessment in estuarine environments.

### 4.4. Microbiota

Exposure to TEB led to rapid and significant changes in the *E. affinis* microbiota, involving a reduction in the alpha diversity and an increase in the relative abundance of the *Gammaproteobacteria* class. Microorganisms are recognized for their sensitivity to environmental changes (Martiny et al., 2023) and members of the *Gammaproteobacteria* have already been reported to be promoted in contaminated environments. These findings suggest the ability of these taxa to resist and/or degrade chemicals, making them potentially valuable bioindicators for evaluating the impacts of contaminants or other stressors (Niepceron et al., 2013, Zhang et al., 2021). Given their vast biochemical potential, microorganisms play crucial roles related to the host's metabolism, immune system, and behavior (Lynch and Hsiao, 2019) and may even influence gene expression in their host organism (Broderick et al., 2014). The microbial changes induced by TEB in this work are therefore likely to indirectly influence copepod fitness and health. In this respect, elevated abundance of the *Gammaproteobacteria* in guts of soil invertebrates were previously found to correlate with oxidative stress or locomotive ability of the host (Zhang et al., 2021). Although technically challenging due to the copepod's size, further studies investigating microbial changes at a more fine-scale resolution than the global approach used here should offer a more complete grasp of changes occurring in the microbiota and their consequences on the host.

## 5. Conclusion

Based on previous results regarding the impact of Tebufenozide (TEB) on gene transcription in *Eurytemora affinis*, we conducted an integrated, multi-scale evaluation of both upstream (epigenetics) and downstream (locomotion and larval development) parameters. Additionally, we examined the indirect effects on its microbiota. While no significant effects were observed on global DNA methylation, TEB was found to impact the larval development, swimming behaviour and microbiota of *E. affinis*, which could affect their fitness and increase their

vulnerability. These results highlight that despite its putative moderate activation of the EcR in *E. affinis*, TEB could still act as a competitor to the moulting hormone at key stages in the life cycle of non-target organisms. The strategy of prioritising the biological responses to be investigated on the basis of gene expression has proven effective. Further processes, particularly those involving the nervous regulation, reproduction, and moulting, should now be investigated to complete this multi-scale approach. Finally, the most relevant responses could be extended to explore potential transgenerational effects. The results show that *E. affinis* can be used to assess a wide range of endpoints on several biological scales. They confirm the relevance of its use as an ecotoxicological model to improve the risk assessment of molecules of concern or environmental matrixes in estuarine environments.

### CRediT authorship contribution statement

**Joëlle Forget-Leray:** Funding acquisition, Conceptualization. **Caroline Arcanjo:** Writing – original draft, Visualization, Validation, Methodology, Investigation, Formal analysis, Data curation. **Céline Boulangé-Lecomte:** Writing – review & editing, Writing – original draft, Visualization, Validation, Supervision, Project administration, Methodology, Investigation, Funding acquisition, Formal analysis, Conceptualization. **Benoit Xuereb:** Writing – review & editing, Validation. **Thierry Berthe:** Funding acquisition, Conceptualization. **Romain Coulaud:** Writing – review & editing, Validation, Investigation. **Anne-Laure Vivant:** Investigation. **Jean-Luc Ravanat:** Visualization, Methodology, Investigation. **Gauthier Trémolet:** Investigation. **Aurélie Duflot:** Investigation. **Yannick Colin:** Writing – original draft, Visualization, Validation, Methodology, Investigation, Formal analysis, Data curation. **Nathalie Giusti-Petrucciani:** Investigation.

### Funding

This research was supported by Region Normandie, grant number 18E01758/18P02520.

### 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

FR CNRS 3730 SCALE

## Appendix A. Supporting information

Supplementary data associated with this article can be found in the online version at doi:10.1016/j.ecoenv.2025.119081.

### Data availability

Data will be made available on request.

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