# Co-contamination with PAHs and TEs reduces the metal uptake of Salix aquatica grandis and increases dissipation

Tristan Couraud, Manhattan Lebrun, Louis Bauchet, Marie-Laure Toussaint, Michel Chalot, Philippe Binet, Lisa Ciadamidaro

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## To cite this version:

Tristan Couraud, Manhattan Lebrun, Louis Bauchet, Marie-Laure Toussaint, Michel Chalot, et al. Co-contamination with PAHs and TEs reduces the metal uptake of Salix aquatica grandis and increases dissipation. Ecotoxicology and Environmental Safety, 2026, 311, 10.1016/j.ecoenv.2026.119839. hal-05499876

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Submitted on 9 Feb 2026

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# Co-contamination with PAHs and TEs reduces the metal uptake of *Salix aquatica grandis* and increases dissipation

Tristan Couraud, Manhattan Lebrun, Louis Bauchet, Marie-Laure Toussaint, Michel Chalot, Philippe Binet, Lisa Ciadamidaro*

Université Marie et Louis Pasteur, CNRS, Chrono-environnement (UMR 6249), Montbéliard F-25200, France

## ARTICLE INFO

Edited by Laura Bulgariu

Keywords:

Soil
Phytomanagement
Mobility
Peroxidase
Willow

## ABSTRACT

The joint presence of various contaminant categories in soil reduces the remediation potential of numerous plant species. However, the mechanisms remain poorly understood, and particularly for tree species. To achieve this goal, we performed a co-contamination experiment with *Salix aquatica grandis* cuttings cultivated for two months under controlled conditions in two soils with two levels (low and high) of trace element (TE) contamination, spiked or not spiked with 600 mg·kg⁻¹ polycyclic aromatic hydrocarbons (PAHs). Our results revealed a significant negative impact of PAH spiking on the mobility of Zn, Cd, and Ni, with average reductions of approximately 35 %. These changes in mobility were associated with lower concentrations in Pb and Cd in plant shoots and a decrease in the total amount of Zn extracted by the plants. These findings emphasize the influence of co-contamination on the phytoremediation potential of high-biomass species. The physiological responses of plants were also affected. Malondialdehyde (MDA) concentrations in leaves increased in response to PAH contamination, indicating oxidative stress, whereas the total antioxidant content varied in relation to TE concentrations in the soil. Furthermore, analysis of the root exudates revealed that peroxidase activity—specifically when the diamonifluorene substrate was used—responded positively to the presence of PAHs, suggesting the activation of enzymatic pathways. Overall, the combination of PAHs and TEs induced diverse and complex effects on both contaminant dynamics and plant physiological health. These results highlight the need to consider co-contamination scenarios in the evaluation and design of phytoremediation strategies particularly for long time protocols.

## 1. Introduction

The co-contamination of soils by trace elements (TEs) and polycyclic aromatic hydrocarbons (PAHs) is one of the most frequent associations due to the ubiquity of these two categories and their joint presence in many anthropogenic activities (Ali et al., 2022). TE contamination in soil may represent approximately 30 % of the 390,000 contaminated wastelands in the European Union (EU), making it the main contaminant category (Panagos et al., 2024), whereas PAHs are ubiquitous byproducts of the incomplete pyrolysis of organic matter and fossil fuels that make them closely related to petroleum-based manufacturing (Alkio et al., 2005). According to the Environmental Protection Agency, co-contaminated soils with both TEs and organic contaminants may represent, for example, up to 40 % of wastelands in the United States (Simón Solá et al., 2019). The presence of different contaminant categories can lead to changes in soil chemical conditions resulting from

chemical interactions between contaminants (Liu et al., 2017). Thus, it is difficult to predict the behaviour of contaminants in the soil; this represents a specific challenge (Xu et al., 2023). It can also increase cytotoxicity, mainly through adsorption processes in the soil, which results in the concentration of xenobiotics near biological membranes (Ali et al., 2022; Khudur et al., 2018; Liang et al., 2016).

Recent works have demonstrated that chemical interactions between TEs and aromatic rings create an electrochemical cation-π interaction that is five times stronger than a hydrogen bond and one-fifth weaker than a covalent bond (Liang et al., 2016; Liu et al., 2017). These bonds reduce both the catabolic activity of certain enzymes and the efficiency of siderophores in capturing TEs due to the proximity of cations and PAHs (Chen et al., 2020; Liu et al., 2017). These interactions also modify the bioavailability of both contaminant categories in the soil, negatively impacting biomass production in certain soil organisms (Jiang et al., 2015). Chemical interactions not only impact the toxicity of elements in

<footnote>* Corresponding author.
E-mail address: lisa.ciadamidaro@univ-fcomte.fr (L. Ciadamidaro).</footnote>
the soil but also pose challenges for remediation strategies, particularly regarding the use of biological approaches.

The remediation of co-contaminated sites is one of the major challenges in mitigating the impact of human activities on ecosystems and preventing the transfer of contaminants through direct exposure or the food chain (Qu et al., 2024). The utilization of microorganisms and plants in remediation processes appears to be a cost-effective, environmentally friendly, and noninvasive method for reducing the impact of contamination (Caracciolo et al., 2024). The bio- and phytoremediation of PAHs and TEs separately are currently well described in the literature, particularly with respect to the biological degradation pathways of PAHs or the sequestration, delocalization, or accumulation processes of TEs in plants (Asemoloye et al., 2019; Imam et al., 2022; Sakshi and Haritash, 2020). Phytoremediation, and moreover phytomanagement, by the valorisation of the produced biomass, could allow to reintegrate contaminated sites in an economical network (O'Hogain and McCarton, 2018). Indeed, fibres and woods could be used, for example, in production materials (Viotti et al., 2022) or for bioenergy production (Chalot et al., 2012). However, the impacts of co-contamination on remediation processes are still not fully understood. There are conflicting results, especially concerning the impact of TEs on the degradation of PAHs (Jiang et al., 2015; Wang et al., 2020). These findings raise questions regarding the impact of co-contaminants on the structure and remediation capacities of plants and soil microbiological communities (Jiang et al., 2015; Liu et al., 2017).

Despite the frequent descriptions of the effects of co-contamination by PAHs and TEs on the biodegradation capacity of microorganisms reviewed by Sandrin and Maier (2003) and Olaniran et al. (2013), few data are available concerning their impact on metallophyte species, particularly in the context of phytoextraction. Indeed, although phytoaccumulation strategies for TEs are often evaluated and optimized through the use of fertilizers or agronomic techniques, less attention has been given to their effectiveness in the presence of organic contaminants (Cao et al., 2022; Mahar et al., 2016). The detrimental impact of organic contamination on the plant metal uptake has already been described (Deng et al., 2018). Indeed, even if numerous studies present the impact of co-contamination on herbaceous or on plants of interest for human consumption (Chigbo and Batty, 2014; Wang et al., 2024; Zhang et al., 2025, 2013), the impact of TE/PAH exposition on woody species is rarely studied, particularly for species with TE extraction capabilities.

Salix species are fast-growing trees capable of withstanding high concentrations of TEs in soil (Evangelou et al., 2015). Although Salix spp. are not considered TE hyperaccumulators, their easily harvestable aboveground tissues, combined with their high yield per hectare, make them relevant candidates for phytoextraction strategies (Courchesne et al., 2017; Kacálková et al., 2015). Some species of Salicaceae have also shown the ability to tolerate and stimulate the degradation of certain organic contaminants, such as pentachlorophenol and PAHs (Marmiroli et al., 2011; Vervaeke et al., 2003). Among the species belonging to the Salix genus, Salix aquatica grandis has been reported to have the capacity to accumulate TEs in aboveground parts, exceeding 2500 mg.kg⁻¹ for Zn and 120 mg.kg⁻¹ for Cd (Ciadamidaro et al., 2019). Combined with different management practices (amendments, co-cropping), the phytoextraction capacities have been shown to be maintained, identifying this species as a good candidate for phytoremediation (Collot et al., 2025). Altogether, these findings identify S. aquatica grandis as a promising candidate for phytoremediation, although its response to co-contamination with PAHs has not yet been investigated.

In this study, we hypothesized that TE/PAH co-contamination might negatively impact the TE accumulation ability and PAH dissipation of S. aquatica grandis. In this regard, the innovative goal of this work was to investigate the impacts of co-contamination on the interactions between S. aquatica grandis and contaminants. To this end, cuttings were cultivated under controlled conditions with TE-contaminated soil spiked with or without PAHs. We hypothesize that (i) TEs are considered

structural parameters of PAH dissipation in soil, (ii) co-contamination impacts S. aquatica grandis health, and (iii) co-contamination decreases the mobility and bioavailability of TEs in the soil and particularly affects the accumulation and phytoextraction potential of S. aquatica grandis.

## 2. Materials and methods

### 2.1. Study site

The soil used for this experiment came from an industrial wasteland located in the city of Vieux-Charmont (47°31'15.8"N, 6°50'23.8"E), France. The soil is heterogeneously contaminated with a wide range of TEs (e.g., Zn. Ni. Pb. Cd. As. Hg) as a result of the site's former use as an uncontrolled landfill for industrial and household waste (Collot et al., 2023). This site is dedicated to the development of strategies of phyto-management that can be replicated at numerous (co)contaminated sites in the regional industrial area.

As part of recent restoration efforts, soil contaminated by organic contaminants (PAHs, polychlorobiphenyls) was removed to allow for the creation of a park open to the public. The main contaminated area is protected to prevent the risk of contaminant transfer to the local population.

### 2.2. Soil sampling and spiking

To assess the impact of co-contamination by TEs and PAHs in the soil of the Vieux-Charmont brownfield site, soil was collected from two areas with either moderate or extremely high TE contamination (Collot et al., 2023). The soil was collected from the top layers (0–30 cm) of the two contrasting locations after the litter was removed in July 2023. Soil harvesting was performed according to the Land Use/Cover Area frame Sampling protocol (Tóth et al., 2013) by mixing soil from different points of the respective area. The soils were then dried at room temperature and sieved to 2 mm to remove the coarsest elements. Each soil sample was then divided into two parts, one spiked with PAH and the other not spiked with PAH. Thus, 200 mg.kg⁻¹ phenanthrene (PHE), fluoranthene (FLT) or pyrene (PYR) was spiked into the soil according to the protocols of Binet et al. (2000) and Madrid et al. (2022) to reach a total of 600 mg.kg⁻¹ PAH in the soil. The required quantity of PAHs was dissolved in 60 mL of dichloromethane and mixed with 10 % of the total soil. After one night and complete evaporation of the CH₂Cl₂, the spiked soil was mixed with the remaining 90 % of the soil. Unspiked soils followed the same contamination protocol with only 60 mL of CH₂Cl₂. The PAH concentrations were determined according to preliminary experimentation during which, Salix aquatica grandis cuttings were exposed to increasing concentration of PAHs (from 0 to 1000 mg.kg⁻¹). The 600 mg.kg⁻¹ appears to be the best compromise between the induction of physiological responses and the limitation of detrimental effects affecting the growth and biomass production. Moreover, the value of 200 mg. kg⁻¹ per compound is consistent with the soil concentrations found in some contaminated sediments (Cornelissen et al., 1998).

### 2.3. Experimental design

In accordance with the soil preparation, four soil conditions were used, and the two contrasted TE-contaminated soils from the wasteland were divided into two conditions as a function of PAH contamination:

* lowest TE-contaminated soil without PAH (L)

* lowest TE-contaminated soil with PAH spiking (L + PAH)

* highest TE-contaminated soil without PAH (H)

* highest TE-contaminated soil with PAH spiking (H + PAH).

For each condition, eleven pots with a 1 L capacity were filled with 750 g of dry soil. Eight pots were planted with 25-day-old Salix aquatica
grandis cuttings (SOUPE nursery, Châtillon sur Chalaronne, France). The cuttings were prepared from cuttings 1.5 cm in diameter and were pregrown in perlite under the same conditions as those used in the experiment. Three pots were left unplanted. To prevent the adsorption of PAHs onto the plastic, the surface of the pots was covered with aluminium foil, and the soil was topped with 150 g of sand (Neptus 0.7–1.3 mm) to prevent volatilization and transfer of PAHs.

The pots were then placed in a phytotron for 58 days (Helios 600, Cryotec, Lunel, France) under the following conditions: day/night, 16 h/ 8 h; 20 °C/16 °C; 70 % humidity; and 74.85 µmol.m⁻².s⁻¹ light for 58 days. The soil was maintained at 70 % of its water retention capacity three times a week with dH₂O.

## 2.4. Soil indicator monitoring

The initial physical (texture, slacking index) and chemical (organic matter (OM) content; phosphorus, magnesium, potassium, and calcium concentrations; pH; C/N ratio; Cation exchange capacity (CEC)) parameters of the soils before spiking (L and H) were measured by SADEF (Aspach-le-Bas, France).

To monitor the variation in pH and the bioavailable TE concentration during incubation, 5 g of each soil sample (L, L + PAH, H, H + PAH) was sampled every two weeks and dried before analysis. The pH was measured using 0.5 g of dried soil, which was shaken for 30 min with 6.25 mL of 1 M KCl and then allowed to settle for 30 min before measurement using a Beckman Φ340 pH meter (Varley, 1972).

## 2.5. Contaminant quantification

### 2.5.1. TE mobility and bioavailability

The concentration of bioavailable TEs was assessed using the protocol of Houba et al. (2000). A mass of 2.5 g of dry soil was stirred with 25 mL of 0.01 M CaCl₂ for 3 h. The solution was then centrifuged and filtered. To prevent microbial growth, the solution was acidified with 35 % HNO₃ (1 % v/v). The concentration of bioavailable TE is expressed in mg.kg⁻¹. To collect the soil solution, the pots without plants, along with five randomly chosen planted pots, were fitted with a soil moisture sampler (Rhizosphere Research Products, The Netherlands) (porous size of 0.15 µm). The solutions were collected using vacuum aspiration overnight. The TE concentration in soil solution are expressed in µg.L⁻¹. Both CaCl₂ and the soil porous solution were analysed using an inductively coupled plasma-atomic emission spectrometer (ICP-AES) (iCAP Pro XP, Thermo Fisher, Besançon, France). The limits of quantification for both types of solutions and the relative standard deviations (RSDs) of the apparatus are presented for TEs in Supplementary Table S1. Certified metal solution (LCG 6027) was used as a standard to verify the calibration of the instrument. The recovery percentages are also presented in Supplementary Table S1.

$$ \text{Chl a (µg.mL}^{-1}\text{)} = 13.36 * \text{A664} - 5.19 * \text{A649} \quad (1) $$

$$ \text{Chl b (µg.mL}^{-1}\text{)} = 27.43 * \text{A} - 8.12 * \text{A664} \quad (2) $$

$$ \text{Carotenoids (µg.mL}^{-1}\text{)} = (1000 * \text{A} - 2.13 * \text{Chl a} - 97.63 * \text{Chl b}) / 209 \quad (3) $$

$$ \text{MDA (nmol.mL}^{-1}\text{)} = ((\text{A} - \text{B})/(157\,000)) * 10 \exp 6 \quad (4) $$

$$ \text{A} = [\text{A}_{532(+TBA)} - \text{A}_{600(+TBA)}] - [\text{A}_{532(-TBA)} - \text{A}_{600(-TBA)}] \quad (5) $$

$$ \text{B} = [\text{A}_{440 (+TBA)} - \text{A}_{600(+TBA)}] * 0.0571 \quad (6) $$

### 2.5.2. PAH extraction and quantification

Soils for PAH extraction and quantification were sampled at the beginning and end of the experiment and freeze-dried before extraction using a Büchi (Rungis, France) SpeedExtractor E-914 system equipped

with 40 mL stainless steel cells. The extraction procedure followed the methods described by Morin-Crini et al. (2014) and was modified using a solvent mixture (50/50, v/v) of acetone and dichloromethane according to Zhang et al. (2016). The analysis was carried out with gas chromatography (GC trace 1600, Thermo Fischer Scientific, Inc., Waltham, USA) coupled with a mass spectrometer (ISQ 7610 Small Turbo, Thermo Fischer Scientific, Inc., Waltham, USA). The column used was a TG-5MS (Thermo Scientific, Inc., Waltham, USA) (30 m × 0.25 mm L·D and 0.25 µm film thickness) coated with 5 % diphenyl/95 % dimethyl polysiloxane. A volume of 1 µL of the sample was injected in splitless mode. Helium was used as the carrier gas at a rate of 1 mL.min⁻¹. The temperature program was as follows: 70 °C (1 min) to 315 °C at 10 °C. min⁻¹ and 315 °C hold for 1 min. The temperatures of the liner, the transfer line and the ion source were 250, 315 and 250 °C, respectively. The electron ionization mode was used at 70 eV and 50 µA. Time-scheduled selected ion monitoring (SIM) was used with each targeted compound in a 10-sec window. For each molecule, three ions (one for quantification and two for confirmation) were monitored. Atrazine 500 µg. L⁻¹ was used as an internal standard for each analysis.

## 2.6. Plant indicator monitoring

To evaluate the impact of soil on plant growth parameters, stem length and number of leaves still alive (green and attached to the stem) were measured weekly. After 58 days of cultivation, the plants cultivated in the low TE contaminated soils presented the first symptoms of starvation, which signalled the end of the experiment.

### 2.6.1. Biomarker and pigment determination

In the middle of the experiment (28 days) and at the end of the experiment (58 days), the fifth-youngest leaf of each plant was harvested, weighed, and quickly frozen in liquid nitrogen to determine pigment and malondialdehyde (MDA) concentrations according to the protocols detailed by López-Hidalgo et al. (2021). The total antioxidant capacity (TAC) was assessed using a TAC Assay Kit (Merk, Inc., Darmstadt, Germany) according to the manufacturer's instructions.

To extract these biomarkers, 1 mL of cold ethanol at 80 % (4 °C) was added to 5–10 mg of powdered leaves that had been ground with liquid nitrogen and freeze-dried. The samples were kept cold, ground with 2 glass beads, and processed 3 times for 10 s at 30 Hz. A volume of 800 µl of the centrifuged supernatant was diluted in the same volume of cold 80 % ethanol.

To measure the concentrations of chlorophyll (Chl) A, Chl B and carotenoids, the optical densities (ODs) were read at 470, 649, and 664 nm. The calculations are detailed in Eqs. (1)–(3) (López-Hidalgo et al., 2021).

The concentration of MDA was determined using a positive thio-barbituric acid solution (+TBA) (0.5 % TBA (w/v) in 20 % trichloro-acetic acid (TCA) (w/v)) and a negative solution (−TBA) (20 % TCA (w/v)). Two hundred fifty microlitres of plant extract were added to the same volume of positive or negative solution. The tubes were heated to 95 °C for 30 min using a heat block, cooled to room temperature and centrifuged at 10,000 × g for 4 min. The supernatant ODs were read at 440, 532 and 600 nm. MDA equivalents (nmol. mL⁻¹) were calculated using Eqs. (4)–(6) (López-Hidalgo et al., 2021).

### 2.6.2. Root exudate (RE) collection and analysis

After 58 days of growth, the plants were carefully removed from the soil and gently washed to remove all the soil particles. The plants were then placed in 0.5 mM CaSO₄ solution (volume sufficient to cover all the root system) and sonicated for 15 min. Afterwards, they were placed in 50 mL of deionized H₂O (dH₂O) for 6 h under the same growth conditions previously described. The solution was then filtered through a 0.22 µm regenerated cellulose syringe filter. An aliquot was directly used for enzymatic analysis, and another aliquot was concentrated using a speed-vac at 30 °C (Ren et al., 2019; Wen et al., 2021) for the
<figure><img src="image_3.png" /><figcaption>**Fig. 1.** Principal component analysis (PCA) of the concentrations of elements (black) and trace elements (underlined red) in the soil solution. The PCA shows 86.35 % of the variance. The colours represent the treatments, and the points shape the absence or presence of plants in the soil.</figcaption></figure>

quantification of phenol content.

### 2.6.3. Peroxidase activity in RE solutions

Peroxidase activities were measured according to the protocols of Jassey et al. (2012) and Criquet et al. (2000). Thirty-five microlitres of the unconcentrated rhizodeposit mixture was suspended in 135 µL of phosphate-buffered saline (PBS) solution (50 mM, pH 6); guaiacol (0.6 mM, ε(470) = 26.6 mM⁻¹·cm⁻¹), 2,2'-azino-bis(3-ethylbenzothiazolin-6-sulfonic) acid (ABTS) (50 µM, ε(420) = 18.460 mM⁻¹·cm⁻¹) or 2,7-diaminofluorene (DAF) (12.9 µM, ε(600) = 10.228 M⁻¹·cm⁻¹) were used as substrates and diluted in PBS. The oxidation rate was monitored at 470, 420 and 600 nm for 15 min. The enzymatic activities are expressed in µmol of oxidized substrate·min⁻¹·g of dry root⁻¹.

### 2.6.4. Total phenolic compound determination

The concentration of total phenolic compounds (TPCs) in the RE solution was determined according to the protocol of López-Hidalgo et al. (2021) using the Folin–Ciocalteu (F–C) reaction. A volume of 100 µL of concentrated RE solution was mixed for 2 min with 200 µL of 10 % F–C solution. The solution was then incubated for 2 h in the dark with 800 µL of a 700 mM Na₂CO₃ solution before the OD was measured at 720 nm. The TPC concentration was expressed as mM gallic acid equivalent. An eight-point calibration curve was prepared using gallic acid solutions ranging from 0 to 1 mM.

### 2.6.5. TE content in plant above-ground tissues

TE concentrations in plant aboveground tissues (stems and leaves) were determined after careful washing with dH₂O, drying at 40 °C and following a mineralization protocol, starting with initial cold mineralization using 1 mL of 67 % HNO₃. An identical quantity of the same acid and 0.5 mL of H₂O₂ 35 % were added, and the sample was heated for 4 h at 120 °C. Certified tobacco leaves (ICT-OBTL-5, Institute of Nuclear Chemistry and Technology, Poland) were subjected to the same mineralization procedure and were used as an internal standard. The solutions were analysed using the same method described in Section 2.5.1. The percentages of recovery are presented in Supplementary Table S1.

## 2.7. Data analysis

All the statistical analyses were performed in R software (v 4.3.1). To

evaluate the impact of soil and plant types on edaphic parameters, analysis of variance (ANOVA) was performed with a contrast adjustment with the sum to compensate for the difference in replicates for each parameter (different pots with or without plants). If the preconditions were not satisfied (absence of normality of the data and absence of homoscedasticity) for an ANOVA, the nonparametric Kruskal–Wallis test was performed. In order to compare the variation rate or percentages between two values, the difference between the final and initial values has been divided by the initial value. In some calculations, the pondered value between the vegetated and the unvegetated pots has been used in order to respect the respective number of all replicates.

Analysis of covariance (ANCOVAs) were performed on the data monitored each week to assess the impact of time on several parameters. A test with a p value of less than 0.05 was considered significant. To analyse the variation in some parameters over the course of the experiment, tests were conducted on the log2 ratio between tₓ and t₀.

# 3. Results

## 3.1. Soil indicator monitoring

Soil physicochemical indicators were monitored on Day 0 in the soil samples from the two study areas, which allowed us to evaluate the differences between the initial soils (L, H) (Table S2). Various parameters, such as the slacking index, OM content, total N, C/N ratio, and CEC, presented similar values for both soils. Conversely, some factors, such as the K₂O/MgO ratio or the P₂O₄ concentration, were relatively high in the H soil. The soil from the L area presented a low phosphate concentration and a low ratio of the major exchangeable element (K₂O/MgO) due to a low concentration of K₂O.

During the experiment, a significant difference of 0.4 units of pH was measured between the two initial soils (L, H), with no relevant impact (lower than 0.05 unit) of PAHs spiking. No further links have been found between the pH variation and TEs uptakes and with the plant physiological responses.

For some elements, particularly the most abundant soil contaminant, TEs, PAHs spiking had a significant effect on ion mobility in the soil solution, although this effect strongly depended on the element (Fig. 1). The PCA clearly revealed a downwards shift for the H + PAH treatment compared with the H treatment, highlighting the effect of PAHs on
mobility, particularly for the Zn, Ni, and Cd concentrations (Table S4).

At the beginning of the experiment, the Zn, Ni, and Cd concentrations were greater in the soil solutions of the H and H + PAH soils than in those of the L and L + PAH soils, particularly for Zn, with average values of 17.8 mg.L⁻¹ in the H treatment and 0.57 mg.L⁻¹ for the L soil, i.e., a factor of 31 times greater. For Ni, the concentrations in the H soil solution were 8.5 times greater than those in the L soil solution (279.1 vs. 33.5 µg.L⁻¹) and 3.2 times greater for Cd (5.93 vs. 1.94 µg.L⁻¹) (values indicate the pondered means between vegetated and unvegetated replicates). For some elements, differences are noticeable at the start between the planted and the unplanted pots (Table S4); however, plant presence did not have significant impacts on the TE concentration in the soil solution for Zn, Ni, and Cd in comparison to the unplanted pots. The impact of PAHs on TE mobility was observed between the H and H + PAH treatments at the beginning of the experiment. PAH spiking was associated with reductions of 41 % (17.80 vs. 10.46 mg.L⁻¹), 39 % (5.93 vs. 3.59 µg.L⁻¹) and 32 % (279.1 vs. 188.2 µg.L⁻¹) (vegetated and unvegetated combined) of the Zn, Cd and Ni concentrations, respectively. After 58 days, the impact of organic contaminants was still significant, but the impact of PAHs was reduced over time: Zn (32 %), Ni (30 %) and Cd (15 %). Even if the effect of PAH on the TE mobility decreased with time, there was no correlation between time and the Zn, Cd, and Ni concentrations in soil solution for the H and H + PAH soil. For the L and L + PAH soil, the concentration of all TE was significantly negatively correlated with the time, indicating a decreasing relationship between time and ion concentrations in the soil solution for these two soil types. Thus, the difference in behaviour was linked to the origin of the soil in the wasteland and not to the presence of organic contaminants. All three main TE contaminants were strongly positively related to each other in the soil solution (Fig. S1). Spearman's correlation coefficients for all relationships (Zn~Ni; Ni~Cd; Zn~Cd) were greater than 0.96.

A different pattern was observed for Sr and Mg, which are two alkaline earth elements clearly more present in the solutions of the L and the L + PAH soil. These two elements were strongly correlated (t = 55.28, p value < 2e⁻¹⁶). In the plant-free soil, the concentrations varied slowly (22 % of the maximum decrease for Sr and 14 % for Mg in the high soil). The presence of the plants was associated with decreases in Mg concentration of 32 % (from 48.71 to 33.10 mg.L⁻¹ for the L and L + PAH treatments) and 36 % (21.90–13.88 mg.L⁻¹ for the H + PAH treatment). For the H treatment, the plant impact was not significant. The presence of PAHs had no impact on the Sr concentration in the L and L + PAH soils, but a decrease was observed in the H + PAH soil compared with the H soil.

Despite a structural effect of PAH contamination on the TE mobility, the behaviour of the four main contaminants found in the wasteland—Cd, Ni, Zn, and Pb—in the bioavailable fraction differed from their respective mobility profiles (Table S5). In unvegetated pots, the presence of PAHs had no observable effect on the TE concentration. In contrast, for the vegetated pots, significant decreases in the bioavailable concentrations of Cd, Ni, and Zn (–20 %, –20 %, and –22 %, respectively) were observed after 14 days of cultivation when the H and H + PAH treatments were compared. For Ni, this effect was no longer significant after Day 14. In the case of Cd, a significant difference was observed only on Day 42, with a 4.7 % reduction in the H + PAH treatment relative to H. The impact of PAHs on Zn bioavailability remained significant between Days 28 and 52, with an 8 % decrease in the H + PAH treatment. The Pb bioavailability did not significantly vary over time. Although the presence of S. aquatica grandis appeared to influence TE bioavailability at specific time points (e.g., Cd on Day 14 in the H + PAH treatment), no consistent trend could be identified to confirm the sustained effect of the plant on TE bioavailability. For the L soil, a decrease in the bioavailable concentrations could be noticed for the unvegetated pots, especially for Mg and P. Overall, the main distinction between TE mobility and bioavailability lies in the persistence of PAH effects in the H + PAH soil: although PAHs had a lasting

<caption>**Table 1**</caption>

Soil concentrations of phenanthrene (PHE), fluoranthene (FLT) and pyrene (PYR) (µg.g⁻¹) as a function of the treatment (L: soil from the lowest TE-contaminated area; H: soil from the highest contaminated area; L + PAH: soil from the lowest contaminated area with PAH spiking; H + PAH: soil from the highest contaminated area with PAH spiking) and the plant presence on soil after 58 days of growth. The values are expressed as the means ± ses. Capital letters are associated with plant replicates, and lowercase letters are associated with no plants. For each PAH, if two conditions share a common letter, the difference is not significant (p > 0.05). (*) indicates a significant difference between the planted and unplanted conditions.

<table><thead><tr><th>Soil type</th><th>Plant in pot</th><th>PHE (µg.g<sup>-1</sup>)</th><th>FLT (µg.g<sup>-1</sup>)</th><th>PYR (µg.g<sup>-1</sup>)</th></tr></thead><tbody><tr><td rowspan="2">L</td><td>NO</td><td>0.93<br>± 0.19</td><td>b<br>0.24<br>± 0.01</td><td>c<br>0.70<br>± 0.24</td><td>b</td></tr><tr><td>YES</td><td>2.14<br>± 0.61</td><td>B*<br>0.26<br>± 0.01</td><td>C<br>2.02<br>± 0.78</td><td>C</td></tr><tr><td rowspan="2">H</td><td>NO</td><td>1.05<br>± 0.03</td><td>b<br>0.76<br>± 0.01</td><td>b<br>1.12<br>± 0.10</td><td>b</td></tr><tr><td>YES</td><td>0.84<br>± 0.08</td><td>B<br>0.57<br>± 0.11</td><td>C<br>0.76<br>± 0.17</td><td>D</td></tr><tr><td rowspan="2">L + PAH</td><td>NO</td><td>14.39<br>± 4.63</td><td>a<br>24.05<br>± 10.55</td><td>a<br>31.91<br>± 11.88</td><td>a</td></tr><tr><td>YES</td><td>15.76<br>± 2.51</td><td>A<br>29.29<br>± 9.28</td><td>A<br>45.04<br>± 12.41</td><td>A</td></tr><tr><td rowspan="2">H + PAH</td><td>NO</td><td>9.31<br>± 0.37</td><td>a<br>16.86<br>± 8.17</td><td>a<br>22.32<br>± 3.49</td><td>a</td></tr><tr><td>YES</td><td>10.77<br>± 0.43</td><td>A<br>11.44<br>± 2.03</td><td>B<br>27.85<br>± 2.37</td><td>B</td></tr><tr><td rowspan="2">Test results<br>(df = 3)</td><td>NO</td><td>χ² = 8.538; p-<br>val = 0.036</td><td>χ² = 9.667; p-<br>val = 0.022</td><td>χ² = 8.641; p-<br>val = 0.034</td><td></td></tr><tr><td>YES<br>(n = 8)</td><td>χ² = 26.75; p-<br>val = 6.64e-6</td><td>χ² = 25.38; p-<br>val = 1.289e-5</td><td>χ² = 25.57; p-<br>val = 1.17e-4</td><td></td></tr></tbody></table>

impact on TE concentrations in soil solution (mobility), their influence on the bioavailable fraction was limited and transient.

## 3.2. PAH dissipation in soil

The recovery percentage of spiked PAH at the initial time of the experiment was 97 % for both soils, but this value varied with concentration. The standard deviation for the H + PAH soil was lower than that for the L + PAH soil. The PHE concentration presented the highest variability and the greatest difference from the expected values. For all spiked molecules, there were no significant differences in PAH concentrations between the L + PAH and H + PAH soils. Owing to the good recovery percentage, the spiking of PAHs was considered satisfactory.

After 58 days, the dissipation rates reached 85 % for the PAH soil and 92 % for the H + PAH soil, with all the molecules combined. The highest dissipation rates were observed for PHE in both spiked soils (Table 1). For all the targeted molecules, significant differences in concentration were detected between the unspiked and spiked soils, despite the high level of dissipation. For the unplanted soil, there was no significant variation between the L + PAH and H + PAH conditions for all the molecules. However, compared with those in the L + PAH treatment, the concentrations in the H + PAH treatment significantly decreased, –8.24 % and –8.9 % for FLT and PYR respectively. Calculations are detailed in the Table S6. These variations were not visible for PHE. After 58 days of growth, plants had no significant effect on PAH concentrations, except for the PHE content in L soil, which was lower in unplanted soil.

## 3.3. Plant indicator monitoring

After 58 days of the experiment, the plants in the H and H + PAH treatments were significantly affected compared with those in the L and L + PAH treatments. Notably, the growth of plants from the H and H + PAH soils strongly decreased from Day 17 onwards, as revealed by stem length and leaf number monitoring (Fig. 2A and B). Between Day 23 and
<figure><img src="image_4.png" /><figcaption>**Fig. 2.** Growth parameters of *S. aquatica grandis*. A: Stem length fold change during the experiment. Stars indicate a significant difference from the L treatment for the corresponding analysis time. B: Leaf number fold change during the experiment as a function of treatment. The stars indicate significant differences with L for the corresponding analysis time. C: Aboveground plant tissue biomass at the end of the experiment as a function of the treatment. If two treatments share the same letter, the difference is not significant (F = 46.357; df = 3; p-val = 3.475e-11). Each point or boxplot represents the mean ± se of 8 replicates.</figcaption></figure>
<figure><img src="image_5.png" /><figcaption>**Fig. 3.** Leaf concentrations of chlorophyll A, B, and carotenoids as a function of the soil treatment after 28 and 58 days of cultivation. The letters represent the results for each pigment and for each measurement step. If two treatments share a common letter, the difference is not significant. Each bar represents the mean ± se of 8 replicates. Test results: Chl A (Day 28: F = 5.93; df = 3; p-val = 0.002***) (Day 58: χ² = 8.92; df = 3; p-val = 0.030*), Chl B (Day 28: F = 5.72; df = 3; p-val = 0.003**) (Day 58: χ² = 9.65; df = 3; p-val = 0.021*), carotenoids (Day 28: F = 1.62; df = 3; p-val = 0.20 ns) (Day 58: F = 4.05; df = 3; p-val = 0.016*).</figcaption></figure>

Day 30, plants exposed to TEs in the H or H + PAH soils lost a significant number of leaves (on average, 4 out of 13 H plants and 3 out of 12H + PAH plants). An ANCOVA of the slope of the fold change from Day 30 to the end of the experiment revealed no significant difference between the treatments (F = 0.237; df = 3; p = 0.87), indicating that the soil type had no impact on leaf production by the plants but did have an impact on their lifespan. *S. aquatica grandis* exposed to H and H + PAH soils presented significant differences in biomass compared with those exposed to L and L + PAH soils (Fig. 2C). *S. aquatica grandis* exposed to H or H + PAH soils produced, on average, 4.2 times less biomass than did

the plants from the L and L + PAH soils, but no effect of PAHs was observed.

### 3.4. Pigment content and biomarkers

Plant health indicators involved in photosynthesis were analysed in seedlings after 28 and 58 days of cultivation in the four soils (Fig. 3). Three pigment concentrations were analysed, i.e., the Chl A and B and the carotenoids. The concentrations of Chl A and Chl B were correlated (S= 2512; p value < 2.2e-16) and followed the same pattern of

<figure><img src="image_6.png" /><figcaption>**Fig. 4.** MDA and TAC concentrations in leaves (nmol.mg⁻¹) as a function of the soil type after 28 and 58 days of experimentation. The letters represent comparisons between treatments for each stage. If two bars share a common letter, the difference is not significant. Each bar represents the mean ± se of 8 replicates. (*) indicates a significant difference between the two-time steps. Test results: MDA (Day 28: F = 2.09; df = 3; p-val = 0.12 ns) (Day 58: F = 8.89; df = 3; p-val = 0.0002***); TAC (Day 28: χ² = 25.54; df = 3; p-val = 1.19e-5) (Day 58: F = 1.97; df = 3; p-val = 0.14 ns).</figcaption></figure>
<figure><img src="image_7.png" /><figcaption>**Fig. 5.** A: Guaiacol (GUA) (mmol of guaiacol oxidized. g-1 fw root. min-1 (green bar)) (F = 8.931; df = 3; p-val = 0.0002) and diaminofluorene oxidase activity (µmol of DAF oxidized. g-1 fw root. min-1 (violet bar)) (F = 6.613; df = 3; p-val = 0.001) as a function of the soil type in the rhizodeposits of S. aquatica grandis. Tests were performed for each parameter separately. B: TPC (mM gallic acid equivalent. g-1 fw root) as a function of the soil type in the rhizodeposits of S. aquatica grandis (χ² = 17.55; df = 3; p-val = 5.4e-4). If two bars share a common letter, the difference is not significant. Each bar represents the mean ± se of 8 replicates.</figcaption></figure>

variation for the two stages of measurement, with the highest content in the plants from the L + PAH soil. After 28 days of cultivation, there was no significant difference in the concentrations of the two Chls between the L and H plants. However, their levels were significantly reduced by 36 % and 33 % between plants from the L + PAH and H + PAH soils for Chl A and Chl B, respectively. After 58 days of growth, the difference was also significant between the L + PAH and H + PAH soils, with reductions of 27 % for Chl A and 23 % for Chl B, indicating that PAH spiking further increased the difference between the H and L soils. The behaviour of leaf carotenoid content was not as sensitive as the chlorophyll contents to the presence of contaminants. After 28 days, no difference was detected between the soils, and after 58 days, the concentration increased in the plants growing on the L + PAH soils compared with those growing on the H + PAH and L soils. The main difference in the Chl contents between the two stages of analysis was the

decrease in the Chl B concentration in the L soil plants; in fact, the Chl concentration in the L + PAH plants significantly decreased compared with that in the L + PAH plants, and the Chl B concentration in the H plants significantly decreased.

The concentration of MDA, a biomarker of lipid peroxidation, was assessed in the leaves (Fig. 4). After 28 days of cultivation, there was no significant difference in the leaf MDA contents among the four soils. At the end of the experiment, significant differences were observed between the soils, with the leaf MDA contents of the H plants significantly increasing by 33 % compared with those of the L plants and by 32 % for the H + PAH plants compared with those of the H plants. At 58 days, the concentration of MDA decreased by 23 % in H plants and increased by 43 % in L + PAH leaves compared with the concentrations on Day 28. The leaf TAC showed significant increases in *S. aquatica grandis* from the H and H + PAH soils on Day 28 compared with that from plants from the

<figure><img src="image_8.png" /><figcaption>**Fig. 6.** Cd, Ni, Pb, and Zn concentrations (µg.g⁻¹) in the leaves and stems of *S. aquatica grandis* as a function of the soil treatment. If two treatments share a common letter, the difference is not significant. (*) indicates a significant difference between organs. Each bar represents the mean ± se of 8 replicates.</figcaption></figure>
<table><caption>**Table 2**
Phytoextraction potential (µg) in aboveground tissues of *S. aquatica grandis* after 58 days of growth on H and H + PAH soils.</caption><thead><tr><th rowspan="2">Soil treatment</th><th colspan="4">Element µg</th></tr><tr><th>Cd</th><th>Ni</th><th>Pb</th><th>Zn</th></tr></thead><tbody><tr><td>H</td><td>2.57 ± 0.30</td><td>6.01 ± 0.50</td><td>3.42 ± 0.50</td><td>958 ± 111</td></tr><tr><td>H + PAH</td><td>1.54 ± 0.20</td><td>6.15 ± 0.50</td><td>1.53 ± 0.60</td><td>662 ± 55</td></tr><tr><td>T-test (df); p-val</td><td>2.83 (10.4); 0.017 *</td><td>0.2 (12.9); 0.846 ns</td><td>2.34 (12.2); 0.037 *</td><td>2.39 (10.1); 0.038 *</td></tr></tbody></table>

L and L + PAH soils (Fig. 4). The TAC contents in the leaves of the H and H + PAH plants increased 2.04- and 2.39-fold, respectively, after 28 days. After 58 days, the difference disappeared between the treatments.

### 3.5. REs of S. aquatica grandis

Various biological indicators, particularly different peroxidase activities and TPC, were measured in the *S. aquatica grandis* REs. Among the different peroxidase substrates, ABTS did not oxidize either oxidase or peroxidase activities (data not shown), whereas guaiacol and DAF peroxidases exhibited striking changes in *S. aquatica grandis* REs from various soils (Fig. 5A). REs from the H and H + PAH soils presented significantly lower guaiacol peroxidation capacities than those from the L and L + PAH soils did, reflecting the inhibitory impact of TEs on rhizospheric guaiacol peroxidase. An additional decrease of 87 % was observed in the guaiacol peroxidation capacity of roots from the H + PAH soil compared with that from the H soil, indicating the cumulative negative effects of TEs and PAHs on rhizospheric guaiacol peroxidation capacity.

In contrast, DAF peroxidase activity increased by 50 % between the REs of plants from the L and L + PAH soils and by 77 % between the REs of plants from the H and H + PAH soils, indicating a strong positive effect of PAHs on rhizosphere DAF peroxidase activity (Fig. 5A). In addition, there was a significant difference between the two unspiked soils from the L and H treatments, highlighting the contrasting and inhibitory effects of TEs on DAF peroxidase activities.

The rhizodeposit TPC showed a contrasting pattern compared with the above monitored RE indicators, with synergistic effects of TEs and PAHs (Fig. 5B). The highest TPC concentration was indeed observed for plants from the H + PAH soils, with 39.6-6.0- and 42.2-fold increases compared with those in the L, H and L + PAH soils, respectively. The TPC of L plant REs did not significantly differ from that of plants exposed to TEs and PAHs, whereas the TPC content of H soil was 7-fold greater than that of L + PAH soil.

### 3.6. TE accumulation in aboveground tissues

The TE concentrations in the leaves and stems and the TE uptake in aboveground tissues were measured for the *S. aquatica grandis* cuttings grown on the four soils. For the four main contaminants of the waste-land, the H and H + PAH leaf concentrations were higher than the L and L + PAH. The PAHs contamination generates three types of behavior as shown in the Fig. 6. The first one is the decreasing of accumulation in the PAH-exposed treatment (L + PAH and H + PAH) in comparison to the unexposed ones (L and H). Indeed, the Cd concentrations decreased by 20 % in H + PAH leaves in comparison to the H plants. Similarly, the Zn concentration in L + PAH plant leaves significantly decreased by 16 % in comparison to the L plants. A comparable result was noticed for the Pb leaves concentration that decreased by 67 % in H + PAH plants in comparison to the H plants. The second possible behaviour was the absence of significant difference between treatments exposed to PAH and those not exposed to PAH (H + PAH vs. H and L + PAH vs. L). This pattern was observed for all stem TE concentrations (Fig. 6). Finally, the

third possible pattern induced by PAH contamination was the increase of TE concentrations in comparison to the unspiked soil. The Ni and Mn (data not shown) leaf concentrations followed this pattern with an increase of 39 % and 62 %, respectively, in the H + PAH plants in comparison to the H plants.

The extracted amounts of TEs (Table 2) presented some differences with the leaf and stem concentrations. For the Ni-extracted content, the differences between the H + PAH and H plants were not significant. For the extracted Cd, Pb, and Zn, the presence of PAHs in soil is associated with a significant reduction of 39 %, 60 %, and 31 % in above-ground tissues, respectively.

## 4. Discussion

### 4.1. TEs are a structural parameter of PAH dissipation in soil

After 58 days of cultivation, dissipation rates of FLT and PYR were higher in the co-contaminated soil (H + PAH) than in the mono-contaminated soil (L + PAH). Interactions between soil parameters, PAHs and TEs are well documented (Babu et al., 2019; Saeedi et al., 2020). The prevailing consensus emphasizes the inhibitory role of TEs on the biodegradation of organic contaminants (Qu et al., 2024; Saeedi et al., 2020; Sandrin and Maier, 2003). However, this framework alone does not fully explain the results observed in the present study. Several factors may account for the contrasting behaviour between H + PAH and L + PAH soils, including (i) shifts in soil microbial communities (Sokhn et al., 2001), (ii) TE-induced modifications of enzyme binding sites and conformations, and (iii) intrinsic chemical properties of the PAHs (Sandrin and Maier, 2003).

In the literature, differences in microbial community composition and associated PAH-degrading capacities among soils are among the most extensively explored hypotheses (Ali et al., 2022; Babu et al., 2019; Liu et al., 2017); however, they are unlikely to be the sole explanation. For instance, Liu et al., (2015) demonstrated that low Pb(II) concentrations can positively affect the activity of both extracellular and intracellular enzymes in isolated bacterial strains, thereby enhancing PAH mineralization. Such enzyme activity modulation, driven by interactions between enzymes and TEs, may redirect enzymatic pathways not initially involved in PAH degradation. A similar mechanism could occur in highly contaminated soils, potentially explaining the increased dissipation of FLT and PYR in the H + PAH soil; however, PHE dissipation did not differ significantly between the two spiked soils.

Finally, the absence of a significant plant effect on PAH dissipation rates suggests that *S. aquatica grandis* does not play a substantial role in the attenuation of this class of contaminants under the conditions tested.

### 4.2. Co-contamination impacted the plant health indicators of S. aquatica grandis

The various plant health indicators monitored in this experiment showed important model variation depending on the parameter considered. Biomass production and stem length of *S. aquatica grandis* were influenced solely by TE contamination. In the literature, the effects of co-contamination on plant growth indicators have already been evaluated, yielding contrasting results. For instance, the negative impacts of 5 and 10 mg.kg⁻¹ Cd were mitigated by the presence of 50 mg. kg⁻¹ PHE and 50 mg.kg⁻¹ PYR in soil. However, at the highest Cd concentration (20 mg.kg⁻¹), PAHs no longer attenuated the toxic effects of TE on biomass production in *Juncus subsecundus* (Zhang et al., 2012). Similarly, in *Festuca arundinacea*, the addition of PYR to soils containing various concentrations of Cu, Cd, and Pb was consistently associated with increased plant biomass (Lu et al., 2014). The presence of penta-chlorophenol, another organic contaminant, produced a comparable effect on the biomass production of *Phragmis australis* exposed to Cd (Hechmi et al., 2014). Collectively, these studies suggest that TE toxicity might be alleviated by organic contaminants, potentially through a
reduction in TE bioavailability and that probably the concentration of organic contaminants has a key role in these modifications. Conversely, other studies have reported that TEs/PAHs co-contamination can exacerbate toxic effects compared to mono-contaminated exposures (Batty and Anslow, 2008; Deng et al., 2018; Mathur and Panwar, 2024).

Regarding pigment concentrations in leaves, they were affected by TE contamination only during the first month of the experiment, with no apparent impact of PAH presence. Similarly, Shukla et al. (2003) reported that the toxicological effects of TEs on plant physiology involve the generation of reactive oxygen species (ROS), which can interact with various biological molecules, including photosynthetic pigments, an effect that was also detected during the first half of our experiment.

In contrast, PAH exposure appears to influence chlorophyll a and b contents. In agreement with our results, this effect has previously been reported in *Arabidopsis thaliana* by Liu et al. (2009). A similar pattern was observed for TAC content in leaves, for which contaminants play a key role in determining its levels. Previous studies have demonstrated a relationship between TAC and the concentration of elements in the soil (Zeneli et al., 2013). Our TAC results are also consistent with findings obtained in S. aquatica grandis, where TAC is considered a marker of TE contamination and toxicity (Collot et al., 2025). The production of stress markers induced by TE contamination may explain the leaf loss observed in the H and H + PAH plants during the first 28 days. After 58 days of cultivation, the leaf markers associated with TE toxicity (Chl A and B, TAC) were no longer overexpressed, which may reflect an acclimation response to the toxic effects of TEs.

The leaf carotenoid content did not significantly vary according to the degree of soil contamination. According to our findings, a previous experiment on *S. aquatica grandis* reported similar results despite an important increase in TE bioavailability (Collot et al., 2025).

Plant stress biomarkers, particularly malondialdehyde (MDA) content, were measured in leaves as indicators of lipid peroxidation and responses to severe oxidative stress (Yuan et al., 2024). Initial analyses after 28 days did not allow discrimination between the different soil conditions; however, after two months of growth, PAH contamination clearly increased MDA concentrations through the generation of reactive molecules, which consequently enhanced membrane damage (Alkio et al., 2005). These results are supported by other studies in which experiments conducted with different PAH concentrations (Song et al., 2011) and in co-contaminated soils (Sun et al., 2011) demonstrated the impact of PAHs on leaf MDA content.

Various types of stress are well known to be associated with increased concentrations of phenolic compounds in the root exudates of many plant species (Michalak, 2006). Phenolic compounds are involved in detoxification processes that reduce the impact of stress. Under TE contamination, these compounds play a key role in the rhizosphere by chelating ions, thereby increasing metal tolerance and protecting plants against TE-induced oxidative stress (Bali et al., 2020). Total phenolic content (TPC) was the only monitored plant health indicator that showed synergistic effects under PAH co-contamination, while exhibiting no response to PAHs alone. It can therefore be hypothesized that the stress response to TEs is enhanced by PAHs, which are also known to generate oxidative stress in plant tissues.

Although PAHs alone were not associated with an increased overall stress response, the activities of some enzymes involved in the oxidation of these compounds increased. Peroxidases are among the main enzymes produced by plants in response to environmental stress and contamination (Gramss et al., 1999; Muratova et al., 2009). The use of various substrates can provide information on the biological origin of these enzymes: ABTS peroxidase, which could not be detected in this experiment, is closely associated with a fungal origin, whereas DAF peroxidase is known to be a sensitive substrate for plant-derived peroxidases (Criquet et al., 2001, 2000; Floch et al., 2007)

Guaiacol peroxidase is closely linked to ROS detoxification pathways and has been extensively studied. The rate of guaiacol oxidation depends on soil origin and is generally known to increase under TE

contamination (Sytar et al., 2013); co-contamination is also reported to have a synergistic negative effect on this enzymatic activity (Babu et al., 2001). However, most studies on guaiacol peroxidase have focused on plant leaves. The discrepancy between our results and those in the literature may therefore be related to the type of root exudates analysed. The release of this peroxidase into the soil may serve functions other than ROS detoxification, although its roles in the rhizosphere remain poorly understood. To explain the decrease in guaiacol peroxidase activity observed in H- and H + PAH-contaminated soils, one hypothesis was proposed by Sat (2008), who studied purified peroxidase from Helianthus tuberosus (L.). This study demonstrated the negative effects of various TEs on enzyme efficiency through binding to the substrate or interaction with the catalytic centre. In our case, high levels of TE contamination in the soil could therefore explain the observed decrease in guaiacol peroxidase activity in the root exudates of *S. aquatica grandis*.

In contrast, DAF peroxidase clearly responded to PAH contamination. Plant-excreted peroxidases have already been implicated in PAH degradation processes in soils contaminated with phenanthrene in studies involving *Medicago sativa* (Dubrovskaya et al., 2017) and *Sorghum bicolor* (L.) (Muratova et al., 2009). According to our results, even though peroxidases are released in the absence of contamination, the presence of PAHs in the soil was associated with increased activities of certain enzymes that may be involved in PAH degradation by *S. aquatica grandis*.

### 4.3. Co-contamination decreased the mobility and modified the accumulation behaviour of TEs in plants

PAH spiking clearly affected the mobility of several elements, particularly Zn, Cd, and Ni. Soil pH is well known to be a major factor controlling element concentrations in soil solution (Carrillo-González et al., 2006). Throughout the experiment, soil pH varied only slightly between the H and H + PAH treatments, thereby ruling out the impact of pH on the decreases in mobility of Cd, Ni and Zn. Changes in both TE mobility and bioavailability were observed in unplanted soils. These variations may be attributed to the biological activity of soil microorganisms that play a key role in TE speciation in soils (Ledin, 2000; Raklami et al., 2022).

To explain the differences observed between planted and unplanted pots at the beginning of the experiment, we hypothesize an effect of the cuttings, particularly their roots, which may alter water flow dynamics within the pots and, consequently, modify ion concentrations in the soil solution (Xiao et al., 2024). After one month, this initial effect appears to be outweighed by other factors, and plant presence no longer resulted in significant differences in TE concentrations in the soil solution. One of the main hypotheses accounting for the reduced concentrations of certain TEs in soil solution involves interactions between soil parameters, such as dissolved organic matter (DOM) concentration and clay content, organic contaminants and TEs. Numerous types of chemical bonds can form between TEs and organic molecules; among the non-covalent interactions, cation-π interactions has been shown to play a significant role (Gao et al., 2006; Li et al., 2024; Liang et al., 2016). These chemical interactions may bind the TE ions dissolved in the soil pore water to PAHs, which are themselves bound to hydrophobic soil particles, thereby reducing TE mobility. The absence of a significant PAH effect on TE bioavailability as measured by CaCl₂ extraction corroborates this hypothesis. Indeed, Ca²⁺ ions used in the extraction procedure may displace TEs complexed with aromatic rings, thereby masking the influence of PAHs. These PAH-induced changes in TE mobility and availability are linked with the accumulation and phytoextraction patterns observed in the aboveground tissues of *S. aquatica grandis*. For Cd and Pb, co-contamination resulted in lower concentrations in leaves and reduced translocation to above-ground tissues. For Zn, PAH spiking affected only the total amount accumulated. Overall, our results indicate that TE mobility is a more reliable indicator of the phytoextraction efficiency of *S. aquatica grandis* than CaCl₂-extractable
<caption>**Table 3**</caption>

Bibliographic synthesis of co-contamination experiment carried on spiked soil. The main results concerning the growth and the TE accumulation are discussed regarding of our own data.

<table><thead><tr><th>Ref.</th><th>Biological model</th><th>Experimental design</th><th>Main comparable result</th><th>Discussion with our data</th></tr></thead><tbody><tr><td>(Batty and Anslow, 2008)</td><td>Brassica juncea and Festuca arundinacea</td><td>Clean soil spiked with PYR at 1000 mg·kg<sup>-1</sup> with or without ZnO at 8000 mg·kg<sup>-1</sup>.</td><td>Co-contamination inhibits B. juncea shoot biomass more than single contamination. For B. juncea, PYR significantly increases the Zn concentration in aboveground tissues.</td><td>Despite a clear difference in Zn accumulation in S. aquatica grandis aboveground tissues, the authors hypothesize that PYR-induced membrane perturbation could be a cause of the higher Zn concentration in B. juncea shoots. Zn spiking could also affect Zn bioavailability and may explain the differences observed with S. aquatica grandis accumulation.</td></tr><tr><td>(K. Wang et al., 2012)</td><td>Sedum alfredii</td><td>Clean soil spiked with 6 mg·kg<sup>-1</sup> Cd (NOs)<sub>2</sub> with or without 25 or 150 mg·kg<sup>-1</sup> PHE or PYR. The tow PAH were not mixed.</td><td>Cd was the only factor influencing the increase in S. alfredii biomass. Both PHE and PYR decreased Cd concentration in shoots at the highest Cd exposure, with PYR showing the strongest inhibitory effect.</td><td>The biomass increase with Cd could be linked to Cd (NOs)<sub>2</sub> providing additional nitrogen supporting plant growth. The decrease in Cd accumulation is consistent with our data. The authors hypothesize a phytotoxic effect to explain the decrease, coherent with the stronger inhibitory effect of the most toxic PAH. They also evoke the higher hydrophobicity of PYR relative to PHE. This hypothesis can be examined with our data, suggesting a PAHs retention effect in the soil solution that decreases TEs availability and consequently plant uptake.</td></tr><tr><td>(Chigbo et al., 2013)</td><td>Brassica juncea</td><td>Clean soil spiked with PYR at 250 or 500 mg·kg<sup>-1</sup> with or without spiked CuSO<sub>4</sub> at 50 or 100 mg·kg<sup>-1</sup>.</td><td>Cu alone did not impact biomass at either concentration. PYR co-contamination significantly affected plant growth at both Cu levels. PYR significantly increased Cu concentrations in aboveground tissues at high Cu exposure for both PYR concentrations.</td><td>The authors hypothesize that increased Cu concentrations in shoots under PYR co-contamination may be linked to PAHs-induced membrane permeabilization facilitating Cu uptake and translocation. Metal spiking versus natural soil contamination could explain differences observed for Cd, Pb, and Zn in our data. TEs are likely less mobilizable in natural soils, and PAH retention effects may outweigh membrane permeabilization effects for certain TEs.</td></tr><tr><td>(Chigbo and Batty, 2014)</td><td>Zea mays L.</td><td>Clean soil spiked with BaP at 1, 5, and 10 mg·kg<sup>-1</sup> with or without K<sub>2</sub>CrO<sub>7</sub> at 50 or 100 mg·kg<sup>-1</sup>.</td><td>Neither contamination nor co-contamination significantly reduced biomass. At the lowest BaP concentration, co-contamination increased Cr accumulation in shoots. At the highest BaP and Cr concentrations, total Cr uptake per plant was reduced. Cr co-contamination increased BaP dissipation at high concentration.</td><td>The authors hypothesize that increased Cr in shoots results from cellular lesions caused by contaminants. They also suggest that reduced Cr availability reduced the metal precipitation and retention in roots, increasing delocalization to aboveground tissues. The increase of Cr under co-contamination is only consistent with the increase of Ni in S. aquatica grandis leaves and supports hypotheses regarding mechanisms involved in this phenomenon.</td></tr><tr><td>(Lu et al., 2014)</td><td>Festuca arundinacea</td><td>Clean soil spiked with PYR at 100 mg·kg<sup>-1</sup> with or without Cu (100 or 500 mg·kg<sup>-1</sup>), Cd (100 or 500 mg·kg<sup>-1</sup>), or Pb (500 or 2000 mg·kg<sup>-1</sup>)</td><td>For Cd and Cu, PYR addition increased biomass. PYR did not mitigate or attenuate the Pb effect. PYR significantly increased TE concentrations in shoots.</td><td>The authors hypothesize that increased TE uptake in the presence of PYR is linked to root membrane permeabilization facilitating penetration of TEs or TE complexes.</td></tr><tr><td>(Jeelani et al., 2017)</td><td>Acorus calamus</td><td>Clean soil spiked with PHE at 50 or 100 mg·kg<sup>-1</sup>, and PYR at 25 or 50 mg·kg<sup>-1</sup>, with or without CdCl<sub>2</sub> at 10 or 20 mg·kg<sup>-1</sup>.</td><td>At the lowest Cd concentration, accumulation was significantly increased by the lowest PHE and PYR exposure. At the highest Cd concentration, accumulation was enhanced by PHE and PYR co-contamination.</td><td>PAH co-contamination produced either positive or non-significant effects on Cd accumulation in A. calamus shoots. These results are consistent with the increase in Ni accumulation in S. aquatica grandis leaves and contradict Cd behavior. The authors highlighted both external and internal effects to explain plant response variability under co-contamination.</td></tr><tr><td>(Zhang et al., 2019)</td><td>Agropyron cristatum L.</td><td>Clean soil spiked with 230 mg·kg<sup>-1</sup> PYR with or without 110 mg·kg<sup>-1</sup> NiCl<sub>2</sub>.</td><td>No impact was reported on Ni concentrations in shoots, but PYR significantly affected Ni concentrations in roots. Despite limited changes in total accumulation, a larger proportion of Ni was bound to cell walls under co-contamination.</td><td>Despite no significant change in aboveground concentrations, changes in Ni binding sites indicate modified plant behavior under co-contamination. This species is not considered a hyperaccumulator, unlike S. aquatica grandis, which may explain differences in Ni accumulation behavior.</td></tr><tr><td>(Liao et al., 2021)</td><td>Pteris vittata L.</td><td>Hydroponic solution spiked with NaAsO<sub>2</sub> (As(III)) or Na<sub>2</sub>HAsO<sub>4</sub>·7H<sub>2</sub>O (V) at 10 mg·L<sup>-1</sup> with or without BaP or fluorene (FLU) (2 mg·L<sup>-1</sup>).</td><td>Both FLU and BaP reduced As(III) and As(V) concentrations in rachises and pinnae, with a stronger effect for BaP.</td><td>The reduction in As uptake is consistent with the decrease of Cd, Pb, and Zn in S. aquatica grandis leaves. The authors explain uptake inhibition by competition between PAHs and As at the root uptake step.</td></tr><tr><td>(Mathur and Panwar, 2024)</td><td>Medicago sativa</td><td>Clean soil spiked with ZnO<sub>4</sub>, PbNO<sub>3</sub> and CdCl<sub>2</sub> at 150 mg·kg<sup>-1</sup> with or without PYR at 500 mg·kg<sup>-1</sup>.</td><td>Co-contamination was more toxic than single contamination but both impacted the plant growth. PYR reduced metal uptake in roots and shoots; TEs decreased PYR dissipation.</td><td>Our data are coherent with this study for:<ul><li>The impact of contamination on plant growth</li><li>The reduction of TE accumulation in roots and shoots.</li></ul></td></tr><tr><td>(Sherri et al., 2024)</td><td>Noccea caerulescens</td><td>Artificially aged TE-contaminated soil (Cu, Cd, Ni, Pb, Zn) spiked or not with 200 mg·kg<sup>-1</sup> PHE.</td><td>PHE significantly affected biomass compared to TE-only exposure. In one population, PHE significantly reduced Zn, Cd, and Ni concentrations in aboveground tissues. Only Cd yield (µg·plant<sup>-1</sup>) was significantly reduced.</td><td>The significant reduction of TE accumulation and uptake is consistent with the reduced Cd, Pb, and Zn accumulation observed in S. aquatica grandis leaves. The authors explain this decrease by organic contaminant accumulation in roots, reducing TEs translocation to aboveground tissues. Opposite patterns were observed for Ni accumulation compared with S. aquatica grandis.</td></tr></tbody></table>

(continued on next page)
<table><caption>**Table 3 (continued)**</caption><thead><tr><th>Ref.</th><th>Biological model</th><th>Experimental design</th><th>Main comparable result</th><th>Discussion with our data</th></tr></thead><tbody><tr><td></td><td></td><td></td><td></td><td>Species specialization—N. caerulescens for Ni vs. S. aquatica grandis for Cd and Zn—could explain differences in TEs delocalization capacity and plant behavior.</td></tr></tbody></table>

Synthesis: Previous research on TEs/PAHs co-contamination in plants, particularly metallophyte species, shows contradictory results. For both plant growth and TEs accumulation, PAHs co-contamination may have either beneficial or detrimental effects. Two main differences distinguish our experiment from previous studies: (i) the natural TEs contamination of the wasteland soil provides more relevant information than metal spiking for understanding plant–TEs–macronutrient interactions; (ii) the biological model is an herbaceous species with TEs-accumulation capacities and a potential human transfer risk through food consumption, whereas co-contamination effects on tree species have not yet been tested.

TE fractions.

The impact of PAHs and other organic contaminants on the TE phytoextraction potential of plants has been reported under a wide range of experimental conditions and for various biological models (Table 3). For instance, reduced TE accumulation was observed in *Brassica chinensis* exposed to co-contamination with a mixture of the 16 priority PAHs (Deng et al., 2018). That study reported significant decreases in Cu, Pb, chromium (Cr), Zn, and Cd concentrations in plants exposed to co-contamination compared with those exposed to TEs alone. At the root level, similar reductions were observed for Cr, Cd and Zn, whereas PAH co-contamination led to increased root concentrations of Pb and Cu.

Our results may be explained by two complementary mechanisms proposed in the aforementioned studies. The first involves chemical interactions in soil between certain TEs and PAHs, which may limit TE mobility and consequently reduce plant uptake (Deng et al., 2018), in agreement with the trends observed in TE mobility in the present study. The second hypothesis suggests that PAHs inhibit TE adsorption and/or translocation through physiological and biochemical disturbances, including alterations in membrane fluidity, changes in electrochemical potential, disruption of cellular ion homeostasis, and inhibition of enzymatic activities such metal-ATPases (Deng et al., 2018; Qu et al., 2024). The increase in lipid peroxidation biomarkers observed in plants in our experiment may support this latter mechanism.

An effect of co-contamination with PHE and PYR was also reported in Cd-spiked soils for the above-ground tissues of *Sedum alfredii* (M.C. Wang et al., 2012). In that study, Cd accumulation decreased with increasing PYR concentrations in soil, in agreement with our findings, which showed a reduction in both Cd concentrations in above-ground tissues and total Cd uptake. In contrast, the effect of PAHs on TE accumulation was opposite for Ni in leaves. An increase in Ni concentration was observed in plants exposed to co-contamination compared with those exposed to TEs alone, whereas PAH spiking did not significantly alter the amount of Ni translocated to aerial tissues. This discrepancy highlights the difficulty of predicting the effects of co-contamination on phytoremediation strategies. Similar increases in shoot concentrations of certain TEs have also been reported under co-contamination with Cr and benzo-α-pyrene (BaP) in *Zea mays* (Chigbo and Batty, 2014). These authors observed enhanced Cr translocation following spiking with 1 mg.kg⁻¹ BaP; however, this effect disappeared at higher BaP doses and under higher Cr concentrations. They attributed this response to toxicological effect of co-contamination, notably increased membrane degradation driven by oxidative stress, which could facilitate the TE entry into plant tissues. Similar results were reported for Cu accumulation in *Brassica juncea* under PYR co-contamination (Chigbo et al., 2013). The authors suggested that increased TE penetration into roots was promoted by PYR diffusion, thereby enhancing TE translocation to above-ground tissues. Collectively, these findings indicate that PAH co-contamination may, under certain conditions, increase the phytoextraction potential of some plant species. For Ni, several studies have demonstrated that Ni–PAH complexes can affect membrane permeability in animal cancer cells (Kim et al., 2019). However, to our knowledge, no relevant data are currently available regarding the effects

of Ni–PAH co-contamination on plant cell membrane, limiting our ability to elucidate the mechanisms underlying the PAH-induced Ni overaccumulation observed in the present study.

From an applied perspective, the results presented here provide valuable insights for the management of co-contaminated sites. PAH/TE co-contamination was shown to reduce the phytoextraction potential of this woody species and may therefore delay the objectives of wasteland restoration. Moreover, our findings highlight the capacity of *S. aquatica grandis* to exude peroxidases involved in the degradation of aromatic compounds, making this species particularly well suited for the management of sites with multiple classes of pollutants.

## 5. Conclusions

The results of this study highlight the influence of co-contamination on the dynamics of TEs and PAHs, as well as on the phytoextraction potential of *S. aquatica grandis*. First, the mobility of the Cd, Zn and Ni in the wasteland soil solution was reduced in the presence of PAHs, suggesting that organic contaminants can modify soil sorption/desorption processes. However, these compounds appeared to affect TE mobility more strongly than their availability, resulting in decreased concentrations in leaves and/or reduced translocation in above-ground tissues. These findings suggest that TE mobility is a more relevant indicator than chemical extractability for assessing the impact of organic contaminants on TE phytoextraction capacity of plants.

Plant health indicators, especially stem growth, biomass production and TAC, were mainly affected by TE contamination, with additional effects associated with PAHs. Nevertheless, a physiological response to PAHs was evidenced by increased plant-exuded peroxidase activity, suggesting the involvement of detoxification or mitigation mechanisms toward aromatic compounds. To our knowledge, this study provides the first assessment of the effects of the PAH/TE co-contamination on the growth and physiology of a woody species such as *S. aquatica grandis*. Further investigations are required to confirm whether similar responses occur in other woody species exposed to multiple categories of contaminants.

Overall, these outcomes emphasize the need to consider the complexity of site contamination when developing phytoextraction strategies. Co-contamination with TEs and PAHs can significantly impair phytoextraction efficiency, not only by inducing physiological damage to plants but also by limiting TE accumulation and translocation within plants. Finally, comprehensive characterization of site contamination may improve phytoextraction strategy by enabling better prediction of TE transfer in soils and their potential impacts on the environment and ecosystem services.

## CRediT authorship contribution statement

**Toussaint Marie-Laure:** Formal analysis, Data curation. **Bauchet Louis:** Formal analysis, Data curation. **Lebrun Manhattan:** Writing – review & editing. **Couraud Tristan:** Writing – original draft, Visualization, Formal analysis, Data curation. **Ciadamidaro Lisa:** Writing – review & editing, Supervision, Project administration, Funding
acquisition, Conceptualization. **Binet Philippe:** Writing – review & editing, Supervision, Conceptualization. **Chalot Michel:** Writing – review & editing, Supervision, Project administration, Funding acquisition, Conceptualization.

## Funding

The project ECOPOLIS has received funding from the Agence Nationale de la Recherche (ANR) under ANR-20-CE22-0010 and from the French General Secretariat for Investment, under grant "Territory of Innovation" of the Investment for the Future Program (PIA3). TC received a grant from the French Ministry of research.

## 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.

## Acknowledgements

We acknowledge Dr. Nadia Morin-Crini, and Caroline Amiot of the PEAT² platform for ICP analyses. We thank Pays de Montbeliard Agglomeration for allowing access to the Vieux-Charmont site. We acknowledge Dr. Aurélie Cébron and Pr. Tanja Mimmo for their help and support during the experiment. We thank also Nolwen Couraud for the graphical abstract illustration.

## Appendix A. Supporting information

Supplementary data associated with this article can be found in the online version at doi:10.1016/j.ecenv.2026.119839.

## Data availability

Data will be made available on request.

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