# Post-exploitation fate of alluvial goldmines in the Amazon: soil erosion, metal(loid) remobilisation and evolution of ecological functions during restoration

Naomi Nitschke, Jennifer Harris, David Amouroux, Emmanuel Tessier, Cédric Legout, Sylvain Campillo, Stéphane Taravella, Stéphane Guedron

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► To cite this version:

Naomi Nitschke, Jennifer Harris, David Amouroux, Emmanuel Tessier, Cédric Legout, et al. Post-exploitation fate of alluvial goldmines in the Amazon: soil erosion, metal(loid) remobilisation and evolution of ecological functions during restoration. Science of the Total Environment, 2025, 996, pp.180150. 10.1016/j.scitotenv.2025.180150. hal-05222596

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# Post-exploitation fate of alluvial goldmines in the Amazon: soil erosion, metal(loid) remobilisation and evolution of ecological functions during restoration

Naomi Nitschke<sup>a,b,c,e,*</sup>, Jennifer Harris<sup>b</sup>, David Amouroux<sup>c</sup>, Emmanuel Tessier<sup>c</sup>, Cédric Legout<sup>d</sup>, Sylvain Campillo<sup>a</sup>, Stéphane Taravella<sup>e</sup>, Stéphane Guedron<sup>a</sup>

<figure><img src="image_4.png" /></figure>

## HIGHLIGHTS

* The restoration (landscaping and replanting) of a goldmine was studied.

* Spontaneous revegetation rapidly decreased particle and metal(loid) fluxes.

* No evidence of the recovery of ecological functions over 5 years after restoration

* A limitation in nitrogen stocks limits the recovery of soil functions.

* Changes in microbial communities on the mining site reflected an adaptation capacity.

## GRAPHICAL ABSTRACT

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

## ARTICLE INFO

**Keywords:**

Gold mining

Amazon

Mine restoration

Soil functions

Mercury

Arsenic

Lead

## ABSTRACT

Artisanal and small-scale gold mining is responsible for deforestation and habitat degradation in Amazonian ecosystems due to soil reworking and increased turbidity in surrounding rivers. In French Guiana, legal mines are required to perform rehabilitation after gold extraction. To assess the successfulness of this rehabilitation, the temporal evolution of particles and metal(loid)s (mercury, lead, arsenic) export was measured, as well as several parameters linked to soil functions (soil carbon and nitrogen, microbial diversity, enzymatic activities). Results show a rapid decrease in particle export (i.e., 3-fold reduction within the first 8 months, and up to 2000-fold reduction after 4 years), linked to an increase in spontaneous colonisation by herbaceous plants. This decrease in particle export induced an efficient decrease in metal(loid)s (Hg, Pb, As) export predominantly associated to the particulate fraction. The parameters associated with ecological functions did not show any significant increase within 5 years after rehabilitation, possibly due to depletion of nitrogen stocks by the

<footnote><sup>a</sup> Univ. Grenoble Alpes, Univ. Savoie Mont Blanc, CNRS, IRD, IFSTTAR, ISTERre, 38000 Grenoble, France</footnote>

<footnote><sup>b</sup> BRGM, F-45071 Orléans, France</footnote>

<footnote><sup>c</sup> Université de Pau et des Pays de l'Adour, E2S UPPA, CNRS, IPREM, Institut des Sciences Analytiques et de Physico-chimie pour l'Environnement et les Matériaux, Pau, France</footnote>

<footnote><sup>d</sup> Université Grenoble Alpes, CNRS, IRD, IGE, 38000 Grenoble, France</footnote>

<footnote><sup>e</sup> SAS GAIA, Cayenne, France</footnote>

<footnote>* Corresponding author at: Univ. Grenoble Alpes, Univ. Savoie Mont Blanc, CNRS, IRD, IFSTTAR, ISTERre, 38000 Grenoble, France.
E-mail address: naomi.nitschke@univ-grenoble-alpes.fr (N. Nitschke).</footnote>
herbaceous growth inducing a competition for resources with soil microbial communities. Microbial communities on the mining site displayed significant evolutions during the 5 years after restoration, with no visible convergence with forest-soil communities, but underlining the adaptability to climatic conditions and a changing environment. These results highlight that current rehabilitation practices are efficient in limiting particle and major contaminants export while more attention needs to be addressed to the recovery of soil functions and biological activities.

# 1. Introduction

Amazon rainforests comprise 50 % of all rainforests worldwide (Xiao et al., 2009). Deforestation has caused forest loss of up to 20 % in the last 40 years, (da Cruz et al., 2020; Fearnside, 2005; Hänggli et al., 2023) leading to losses in biodiversity (Brandão et al., 2022) as well as warming at the local and regional scales (Artaxo, 2023; Butt et al., 2023).

Amazonian soils are typically highly weathered and nutrient-depleted compared to temperate soils (Quesada et al., 2010). The presence of a rainforest above these soils is made possible by rapid and efficient nutrient cycling thanks to fast organic matter (OM) degradation by soil (micro)organisms present in the thin organic top layer, followed by reuptake by trees (Wagg et al., 2014). However, this equilibrium is fragile, and once the thin organic top-layer is lost, the recovery of a forest ecosystem is not guaranteed. Hence, deforestation in the Amazon basin is often linked to a decrease in soil biodiversity (Franco et al., 2019). In addition, deforestation leads to enhanced soil erosion, particularly in tropical regions where rainfall is intense and frequent (Labrière et al., 2015; Lal, 1990). Such a loss in nutrients and microorganism diversity has long-term consequences on soil properties, with impacts observed on soil functions even decades after reforestation (Veldkamp et al., 2020).

At the scale of the Amazon region, the main drivers for deforestation are agriculture and artisanal and small-scale gold mining (ASGM), although regional differences exist (Hänggli et al., 2023). In the Guiana Shield, due to extensive gold mining and less extensive agriculture, ASGM is the major cause of deforestation, mostly in Guyana and Suriname, but also to some extent in French Guiana (Hänggli et al., 2023). In the case of ASGM, erosion is exacerbated, as alluvial gold mining involves not only the removal of the vegetation cover but also topsoil stripping using water hoses in order to access the gold-bearing gravel layer. This leads to important exports of particles to rivers, which negatively affects downstream ecosystems (Mol and Ouboter, 2004; Tudesque et al., 2012). Such an increase in river turbidity is accompanied by an increase in geogenic mercury (Hg) fluxes, which results from the accumulation in the soil of atmospheric Hg taken up by vegetation over millions of years (Fostier et al., 2015; Guédon et al., 2006). Once deposited on the river bed, Hg bound to fine particles can undergo methylation and biomagnification along the food webs, where methylmercury (MeHg) neurotoxicity can cause health problems in indigenous communities relying mainly on fishing (Maury-Brachet et al., 2006; Pignoux et al., 2019). Additionally, the discovery of lead (Pb) poisoning and high blood lead levels in children of French Guiana communities in 2011 (Rimbaud et al., 2017) prompted the creation of a working group supported by the regional health agency to unravel the contamination sources and mechanisms of different heavy metals and metalloids (including arsenic) in French Guiana ("Stratégie Métaux Lourds", StrAMELO). High blood lead levels were proposed to be linked mainly to manioc consumption, and possibly to exposure to lead bullets through hunting activities (Andrieu et al., 2020; Maurice et al., 2021).

Numerous studies have documented the direct impact of ASGM on Hg pollution and SPM release during operations (Dethier et al., 2019; Diringer et al., 2020; Pfeiffer et al., 1989; Telmer et al., 2006). Presently, there is a strong demand from stakeholders and authorities about the fate of ASGM sites after operations and/or rehabilitation with regard to these two aspects (Hellal et al., 2020). After clear-cutting, the ASGM

area is usually first colonised by pioneer species (mainly herbaceous), which are less nutrient-demanding and resist strong insolation. This reduces soil erosion and provides organic matter and nutrients for other plants that appear later on in the succession towards the potential establishment of a secondary forest. Successful rehabilitation therefore attempts to restore soil stratification, stabilise them, and ultimately allow the accumulation of newly produced OM. Therefore, this promotes the stabilisation of geogenic Hg originally present in these soils thanks to its affinity for OM and iron oxy(hydr)oxides, thereby reducing its export to the surrounding watershed. The availability of nutrients and their carrier phases in the soil can be a limiting factor that prevents the creation of a secondary forest, especially carbon and nitrogen (Gomes and Luizão, 2011), but also phosphorous (Cunha et al., 2022), and trace elements, which can be involved in essential functions like nitrogen-fixation (Barron et al., 2008).

In French Guiana, legal gold mines are required to restore the mining site after exploitation. Mines are therefore subject to controls during and after exploitation to limit their impact on the environment and assess their rehabilitation efforts (Marquis, 2015). To promote rapid vegetation recovery in addition to natural regeneration, replanting is performed on at least 30 % of the mining site, usually with plants from the Fabaceae family, which are able to form a symbiosis with nitrogen-fixing bacteria through root nodules. However, no regulations exist for gold mine rehabilitation concerning microbial diversity or soil functions, which are key parameters for trace elements stability and vegetation growth (Fikri et al., 2021; Hellal et al., 2023). Few studies in the area have investigated the evolution of soil functions on a mining site after restoration (Couic et al., 2021; Schimann et al., 2012), and the timespan for the recovery of these soil functions is yet poorly documented. Although it is expected that the re-establishment of vegetation cover will stabilise the soil and limit erosion, important particle fluxes, including exports of metals and organic matter, are therefore expected after landscaping, until the soil is covered by pioneer vegetation.

Hence, the overall objective of this study is to follow the short-term temporal evolution of a gold-mined flat after exploitation and its impact on both contaminant fluxes and on the recovery of soil functions. To this end, soil erosion and metal(loid) contaminant fluxes (Hg, Pb, As...) were monitored on experimental plots to document their evolution during rainfall events over several years after the mining site rehabilitation. In addition, soil biotic and abiotic indicators related to soil functions were analysed (i.e., total carbon and nitrogen contents, soil microbial diversity through DNA metabarcoding and enzymatic activities) to document the evolution of soil functions during the human-assisted mine restoration process.

# 2. Materials and methods

## 2.1. Study area

French Guiana (87,500 km²) is part of the Guiana Shield located in the north-eastern region of South America between Brazil and Surinam (2–6°N and 51°30′–54°30′W). It experiences an equatorial humid climate with an average annual rainfall of around 4000 mm and 96 % of the territory is covered by a tropical rainforest. Samples were collected from the Awa site, an alluvial gold mine (54°20′34″ W, 4°19′8″ N) located on the eponymous river and under control of the mining company SAS GAIA (Fig. 1). The Awa River is part of the bigger Maroni watershed, the
largest river of French Guiana (66,814 km²), which is densely gold-mined (Gallay et al., 2018) and has a long history of gold mining dating back to the 1900s (Jébrak et al., 2021). The Awa watershed area is 32 km².

Alluvial gold mining exploitation was performed below a riverbed, to extract gold from gravel deposits resulting from the weathering of auriferous quartz veins (Thomassin et al., 2017). These downslope valleys typically consist of sandy hydromorphic soils exhibiting reducing conditions due to the presence of water flowing very slowly (Guédron et al., 2009). The mineralogy of these soils is dominated by quartz and kaolinite, with some goethite. After exploitation, restoration is performed on the mining sites. Restoration consists in filling the mining ponds with the gravel and clay removed during exploitation, performing landscaping, and then replanting trees grown on nurseries on-site, at a three-meter interval. The tree species used are *Euterpe oleracea* (açaí palm), *Inga* sp., *Senna multijuga*, and *Hymenaea courbaril*, the three latter being part of the Fabaceae family.

On the Awa mining site, exploitation and production was conducted on the site from downstream to upstream between September 2017 and April 2021. The first restoration works were operated on some areas between May and July 2019. Full restoration was then performed between May 2021 and August 2022.

Four sampling campaigns were conducted over three years (from 2022 to 2024) to follow the evolution of restoration on the mining flat:

G1 (04/22), G2 (12/22), G3 (05/23), G4 (04/24), on two different sites (Fig. 1). One restored and replanted in April 2022, just before the first sampling campaign, and one restored three years before. Hereafter the two sites are termed “landscaped” and “tree-covered”, respectively. The samples were labelled based on the time since restoration, with samples on the landscaped site labelled T = 0 months (G1), T = 8 months (G2), T = 13 months (G3), and T = 24 months (G4) and the samples on the tree-covered site labelled T = 36 months (G1), T = 44 months (G2), T = 49 months (G3), and T = 60 months (G4) (Fig. 1). For soil function analyses, a nearby reference site under forest cover was studied during the last sampling campaign (G4).

## 2.2. Experimental design for erosion monitoring

Erosion plots were installed in triplicate at each sampling site to recover the eroded material from a defined area during a rain event. Each plot consisted of wood panels (2 m × 2 m, 4 m²) inserted 10 cm into the soil and extending 15 cm above the soil. Each plot was connected to an 80 L bucket installed at the lowest point of the slope and buried in the soil.

### 2.2.1. Meteorological data

Hourly temperature and rainfall data for the determination of annual meteorological trends were retrieved from the Meteo France

<figure><img src="image_6.png" /><figcaption>**Fig. 1.** Study site, timeline with rainfall data (Grand-Santi meteorological station (meteo.data.gouv.fr [last accessed on 3rd February 2025])) and sampling points for the erosion (round) and soil function (diamond) experiments.</figcaption></figure>
meteorological station at Grand-Santi (GS) (meteo.data.gouv.fr [last accessed on 3rd February 2025]), which is located 7 km away from the mining site. For specific events, a garden rain gauge was used and the duration of the rainfall event was recorded.

### 2.2.2. Sample collection

All samples were collected in pre-cleaned vials following ultra-clean technique protocols (Stoichev et al., 2006). Surface runoff water and particles were collected after each rain event during the sampling campaigns G1, G2, and G3. The total volume of water was measured, a 250 mL aliquot of the suspension was sampled in an FEP bottle, to measure both SPM, Hg speciation, and major and trace elements in the filtered and particulate fractions. The rest was discarded and the coarse matter sedimented at the bottom was retrieved, freeze-dried and weighed. The 250 mL aliquot was filtered on a 0.45 µm PVDF filter (Sharif et al., 2014; Stoichev et al., 2016). The filtrate was collected in FEP bottles acidified to HCl 0.02 % for Hg analyses, and in trace grade falcon tubes acidified to HNO₃ 2 % for multi-elemental analyses. The PVDF filters were air-dried in a laminar flow hood for 48 h and weighed to determine the amount of particles. This particle mass was divided by the total water volume that was filtered in order to determine the suspended particulate matter (SPM) concentration in mg.L⁻¹.

## 2.3. Experimental design for soil function monitoring

Soils were sampled on the two sites monitored for erosion, i.e., the landscaped site and the tree-covered site during three out of the four campaigns (G1, G3, G4), at one-year intervals. Sampling was always performed during the beginning of the rainy season to document soil ecology in similar seasonal conditions as soil biological activity can be greatly impacted by seasonal changes (Cusack et al., 2019; Madegwa and Uchida, 2021). Additionally, samples were also collected from a forest site at G4 as a reference site for comparison.

### 2.3.1. Sample collection

From each sampling spot, three areas of ~20 m² were defined and 9–10 soil cores from the 0–20 cm depth were sampled using a manual auger. These cores were pooled and homogenised to form a composite sample for each replicate for further physico-chemical and biological analyses. Soil cubes of 10 × 10 × 10 cm were also collected from each site to measure soil density as described below. Plant above- and below-ground biomass was also collected and analysed as described below.

### 2.3.2. Chemical analysis

All details for protocols and QA/QC are provided in S.I.1.2. Briefly, total Hg on solid coarse samples was analysed by atomic absorption spectroscopy after pyrolysis and gold amalgamation (AMA 254, Altec) as described elsewhere (Guédron et al., 2009; Roos-Barraclough et al., 2002). Hg speciation in filtered water and on suspended particulate matter (SPM) collected on filter samples was analysed following established methods using single-spike isotope dilution analysis by gas chromatography – inductively coupled plasma mass spectrometry [SS-ID GC-ICP-MS (Cavalheiro et al., 2016, Monperrus et al., 2008, Monperrus et al., 2005, Sharif et al., 2014)]. The concentration of major elements and trace metal(loid)s (i.e., As, Pb) was measured by inductively coupled plasma-atomic emission spectrometry (ICP-AES) (Aglent, Varian 720-ES) and by inductively coupled plasma mass spectrometry (ICP-MS) (Agilent 7900 coupled to an ESI PrepFast IC module) respectively. Organic carbon (C) and nitrogen (N) concentrations were obtained using a Flash Smart elemental analyser (Thermo-Fisher Scientific) (Piton et al., 2020).

## 2.4. Soil erosion measurements

Soil erosion rates and critical erosion shear-stress were measured using an EROMES device as described elsewhere (Haddad, 2022; Joensuu et al., 2017; Lanuru et al., 2007; Schünemann and Kühl, 1991), and detailed in Section S.I.1.3.

## 2.5. Soil physico-chemical characterisation

Soil Water Holding Capacity (WHC), apparent density (DA), pH, and granulometry were measured as detailed in the supplementary material (S.I.1.4). Briefly, soil density was obtained by collecting soil volumes of 10 × 10 × 10 cm from each site. The total volume of soil removed was measured by filling the hole (covered with a plastic foil) with water. The total amount of soil was weighed, and an aliquot was taken to be dried and weighed again, to measure the soil apparent density (DA). pH was measured in the laboratory after resuspending 5 g of soil in mQ water volume of 25 mL. Soil granulometry was obtained using a laser granulometer (Malvern, Mastersizer 2000) following the protocol detailed in (Grangeon et al., 2012). Finally, plant biomass over the soil surface was obtained by cutting the aerial part of the plants, which were washed with H₂O mQ, and dried. The above-ground biomass (AGB) was then calculated by dividing the plant weight by the total soil surface. Roots were also separated manually, washed with H₂O mQ, dried, and weighed. The root density (RD) was calculated by dividing the weight of dried roots by the total soil volume.

## 2.6. Soil biological analyses

Soil extracellular enzymatic activities of arylsulfatase (ARS), β-glucosidase (β-GLU), N-acetyl-glucosaminidase (NAG), and phosphatase (PHOS) were measured according to ISO 20130:2018. Soils for enzymatic analyses were frozen at -80 °C prior to analysis.

Soil DNA was extracted from 0.6 to 0.7 g of soil (dry weight), previously frozen at -20 °C, using the FastDNA® SPIN Kit for soil (MP Biomedicals, Illkirch, France) and the FastPrep R-24 classic Instrument used per the manufacturer's instructions. Soil DNA extract concentrations were measured with the Quantus fluorimeter and the Quantifluor™ dsDNA System, according to the manufacturer's instructions (Promega, Madison, WI, USA). Real-time quantitative polymerase chain reaction (qPCR) was performed to assess gene abundances as a measure of bacterial abundance (16S rRNA) and abundance of genes linked to the nitrogen cycle (narG, amoA, nosZ, nifH, nirS). The primers used are listed in Table S1.

Bacterial and archaeal diversity were determined by 16S rRNA gene metabarcoding. Sequences were processed on the Galaxy platform using FROGS (Escudie et al., 2018) based on the Galaxy analysis platform (Galaxy, 2022). Bacterial double affiliation was performed by blasting amplicon sequence variants (ASVs) against the SILVA database 138.1 (Quast et al., 2013). ASVs with affiliation <100 % at the phylum level were removed from the dataset.

## 2.7. Statistical analyses

All statistical analyses were performed using R. The Student's t-test was performed when the data was normally distributed and the variances between the two studied groups were equal. When this was not the case, the Wilcoxon test was performed. Regressions are shown only if the p-value is below 0.05. To visualize dissimilarities in bacterial community structures, two-dimensional nonmetric multidimensional scaling (NMDS) based on Bray-Curtis dissimilarities of log-transformed data was
applied using the R software. The relationship between community profiles and hydrogeochemical variables was evaluated by fitting environmental vectors a posteriori onto the NMDS.

### 3. Results and interpretation

#### 3.1. Fate of surface soil after landscaping

##### 3.1.1. Anthroposol characteristics

Reworked soils after mining operations consist mainly of sand and gravel mixed with the fine fractions that settled in the tailing banks. The physicochemical characteristics of the top-soils (0–20 cm) on the mining site are rather consistent over time and differ mostly between the two sites by granulometry, with the landscaped site being sandier (66 ± 2% on the landscaped site and 50 ± 5% on the tree-covered site), and the tree-covered site showing higher proportions of clay and silt (Fig. S4). This increase of the clay fraction over time has also been observed, on a much larger timescale, in the restoration of a sandy deserted land in China, and was attributed to the sequestration and immobilisation of fine particles by the vegetation cover (Zhang et al., 2016). The water holding capacity (WHC) of the soil was similar between the two sites. The total mercury (THg) concentration in these mine top-soils is 102 ± 36 ng.g⁻¹, somewhat lower than what was measured on pristine soils around the mine site (143 ± 45 ng.g⁻¹), which is typical for non-contaminated soils in French Guiana and the Amazon region (Beliveau et al., 2017; Goix et al., 2019; Guédron et al., 2018). The lower Hg concentrations on the mine soil compared to the pristine soils can be attributed to the reworking and mixing of the soil over several meters, where deep soil horizons are depleted in Hg as observed here (42 ± 10 ng.g⁻¹ below 2 m), and elsewhere in French Guiana (Guédron et al., 2006). This can also be due to Hg loss from soils to the river during the mine exploitation phase. Many studies report an increase in Hg concentrations in suspended matter and sediments downstream of gold-mining areas (Lino et al., 2019; Moreno-Brush et al., 2020), but few studies have focused specifically on the export of natural Hg from a mining site during the exploitation phase (Grimaldi et al., 2015). Concerning Pb, soil concentrations were similar in the mined soils and in the forested soils (3.2 ± 1 µg.g⁻¹), but As concentrations were significantly depleted in mine top-soils (0.27 ± 0.04 µg.g⁻¹) compared to pristine soils (1.2 ± 0.4 µg.g⁻¹).

##### 3.1.2. Rapid attenuation of particles and contaminant export from soil

To document the evolution of reworked soils during the mine restoration, exported particles and contaminants were studied in experimental plots from the initial landscaping (T0) to 49 (T49) months. Immediately after landscaping and revegetation (T0), the export of soil particles during a rain event was high (78 ± 48 g.m⁻²), and in the range of values on bare agricultural land in Brazil (40 ± 20 g.m⁻²) (Beliveau et al., 2017), indicating strong erosion of the mining flat right after restoration. In contrast, the amount of eroded soil particles on the tree-covered site (T = 49 months since rehabilitation) was significantly lower (p = 0.001, Wilcoxon), and virtually reduced to zero, at 0.01–0.3 g.m⁻², which is comparable or even lower than what was observed on forested areas in Brazil and other tropical forest soils (Beliveau et al., 2017; Hartanto et al., 2003). Because the amount of eroded particles depends on the rain event intensity, particle exports were normalised by the total amount of rainfall (in mm) in order to derive particle fluxes. The total particle flux exported during a rain event was significantly decreased (p = 0.0034, Wilcoxon) 4 years after rehabilitation, leading to a 2000-fold reduction of particle export when taking into account all eroded particles (fine and coarse), and still a 200-fold reduction when considering only the fine (suspended) particles (Fig. 2A).

Concentrations of total mercury (THg) and its speciation as methylmercury (MeHg) as well as other trace metal(loid)s (i.e., Pb, As) were measured on eroded particles and in the filtered fraction. Mercury concentration on the fine particles was 172 ± 73 ng.g⁻¹, and did not

differ significantly between the two sites, suggesting that the nature of particles exported from the two sites is similar. In the filtered fraction, THg was 1.5 ± 0.8 ng.L⁻¹ (Fig. S1A), with no significant difference between the two sites (p = 0.6), or evolution during the time after rehabilitation.

Methylmercury (MeHg) accounts for less than 5 % of total Hg, but this proportion significantly (p = 0.001) increased over time, likely due to an increase in the particulate MeHg concentration over time (p = 0.014). Particulate Hg dominated the partitioning between particles (HgP) and water (HgF), representing almost 100 % of total Hg (THg = HgP + HgF) at T0, and more than 75 % at T49 (Fig. S1A). As a result, the total Hg fluxes are mainly driven by the erosion of the mining flat soil, and the decrease in particle export observed after 49 months results in a steep decrease in total Hg fluxes (p = 0.003, Fig. 2). However, because the decrease in water export was less extensive (40-fold reduction) than the fine particle export (200-fold reduction) between T0 and T49, the proportion of mercury (Hgp) associated with the particulate fraction significantly decreased over time (Fig. S1A).

Lead (Pb) concentrations in the particulate fraction were stable over time and were identical in the landscaped and tree-covered sites at 12 ± 6 µg.g⁻¹, in the same range that has been measured in French Guianese topsoils (Guédron et al., 2006; Maurice et al., 2021). In contrast, As concentrations were significantly lower (p = 0.00012) in the tree-covered site (1.4 ± 0.4 µg.g⁻¹) than in the landscaped site (2.8 ± 0.8 µg.g⁻¹), in the range of soils in the Brazilian Amazon region (da Silva Junior et al., 2019). In the filtered fraction, both Pb and As concentrations were constant over time after restoration with a mean of 0.08 ± 0.1 µg.L⁻¹ and 0.04 ± 0.02 µg.L⁻¹, respectively (Fig. S1B). The partitioning coefficient between the filtered and particulate fractions ([PbF]/[PbP] or [AsF]/[AsP] using concentrations in µg.L⁻¹) was generally lower (by a factor 10) for Pb than for As. As a result, the particle export reduction over time results in a more drastic decrease of the proportion of As in the particulate fraction (down to 52 ± 13 %) compared to Pb (83 ± 12 %), which exhibits a behaviour that is more similar to Hg. In any case, the reduction of both particle and water export over time leads to a significant reduction of both As (150-fold) and Pb (100-fold) fluxes over time (Fig. 2).

##### 3.1.3. Pioneer plants (herbaceous) as a basis for soil surface stabilisation

Over the first year following landscaping, a sharp increase in root density due to spontaneous colonisation by herbaceous plants was observed (Fig. 2E). In addition to an increase in root density, the above-ground biomass also increased with time, with high variability between plots and sites. Finally, the presence of biological soil crusts was observed on the soils (Fig. S3), resembling a biofilm- or moss-like matrix covering soil surfaces. No ecological characterisation has been performed on these crusts. The herbaceous plants, together with the colonisation of soil surfaces by biological soil crusts are the most likely hypothesis for the observed decrease in particle and contaminant export, as has been observed elsewhere (Loebmann et al., 2020; Oliveira and Maciel-Silva, 2022).

A large heterogeneity in particle and contaminant exports was observed at T = 13 months between the 3 plots, and was linked to a heterogenous plant cover between the tree plots at that time point (Fig. S2). Indeed, these plots were located in proximity to the forest edge, with the plot closest to the edge displaying the strongest herbaceous colonisation. The nutrient inputs from the forest edge may have affected the ability of pioneer plants to colonise the site. Thus, the variability in particles and metal(loid) export reflects the heterogeneity of plant colonisation within a site, which is then smoothed out over time as herbaceous colonisation becomes ubiquitous.

Such observations were confirmed by erodibility experiments performed on surface soils at the edge of each plot at T = 13 months. An increased erosion constant coupled to a lower critical erosion threshold (Fig. S2A) were found on the soil with little herbaceous cover, whereas the plot with more pronounced vegetation showed an increased soil
<figure><img src="image_7.png" /><figcaption>**Fig. 2.** Evolution of exports of A) particle, B) Hg, C) Pb, D) As normalised by rainfall (mm) and E) root density over time after rehabilitation.</figcaption></figure>
stability. This further characterises the role played by the spontaneous colonisation of herbaceous plants in the stabilisation of soil and contaminants on the mining site. These erodibility measurements cannot be directly related to the particle flux observed during the rain events, as the experimental setup is designed to simulate shear stress due to surface runoff and disregards the effect of raindrops on soil aggregate breakdown (Legout et al., 2005), but it provides an indication of differences in erodibility between the sites.

Hence, the rapid decrease in particle and contaminant export after 1 year of rehabilitation of the mining flat is mostly attributed to the spontaneous colonisation of herbaceous plants that helps to stabilise the soil. However, in order to go beyond the vegetative state and obtain a closure of the mining site, with a secondary forest, several functional cycles in the soil need to be restored. These functional cycles (e.g. carbon and nitrogen cycles) involve the action of microorganisms. Microbiological indicators were thus followed over time to assess the recovery of soil functions on the restored mining site.

## 3.2. Evolution of soil functions and microbial communities during revegetation

### 3.2.1. Microbial biomass and carbon and nitrogen cycles

To study the recovery of soil functions, the six plots ("landscaped" and "tree-covered" triplicates) were compared to an undisturbed reference forest site located close to the mining site.

Analysis of biological indicators, i.e., total C and N content in the soil, total DNA and 16S qPCR as proxies of microbial biomass, and functional genes linked to the N-cycle (narG, nifH, nirS, nosZ) did not show any significant increase over time after restoration (Fig. 3).

Total C and N contents are much lower for the mining site than for the nearby forest (p = 0.001 and 0.008 respectively, Fig. 3), and in the range of data reported for other mining sites in French Guiana (Couic et al., 2021; Schimann et al., 2012). This was expected as most of the organic layer is usually eroded and washed away during the exploitation and early restoration phases. On the forested site, in addition to the

higher concentrations of C and N, higher amounts of total DNA and functional genes linked to the N-cycle were detected. C and N trends on the mining site showed a decrease over time, with a sharp decrease after 13 months, followed by low C and N contents afterwards (Fig. 3). The C/N ratio was significantly (p = 0.0002) lower on the tree-covered site (11.1 ± 1.4) than on the landscaped site (14.3 ± 0.8), due to a steeper decrease in C than in N over time. Landscaped site C/N ratio is close to the one for the reference forest site at 13.7 ± 0.1, which is also similar to reported values in forests in French Guiana (Schimann et al., 2012).

Amongst indicators of soil microbial activity, only arylsulfatase, an enzyme involved in the sulfur cycle, which hydrolyses sulfuric esters to phenols and sulfates, displayed an increase in activity over time (ARS, p = 0.002, linear regression, Fig. S5). This suggests that ARS might be a relevant early-stage indicator to detect precursors of soil microbial activity recovery.

### 3.2.2. Evolution of microbial communities on a restored gold mined flat

From a total of 1551 ASVs obtained from the 16S rRNA sequencing, the major bacterial phyla observed in soil samples (Fig. S8) were Acidobacteria (33 ± 6 %), Proteobacteria (27 ± 6 %), Chloroflexi (16 ± 5 %), Actinobacteria (8 ± 2 %) and Planctomycetia (5 ± 1 %), with relative abundances in the range of what was previously found in tropical forest soils in French Guiana (Zinger et al., 2019), and on gold mining sites in Colombia (Cardona et al., 2024). Alpha diversity indices (Shannon, Chao1, inverse Simpson) did not show any significant differences in diversity between mine soils and the forest soil. However, a difference in alpha diversity was observed only for the time point T49 compared to all the other points before and after. This indicated a reduced richness (Chao1), but an increased community diversity (inverse Simpson index) compared to other mine soils and the forest (p = 0.00164, ANOVA).

Non-metric multidimensional scaling (NMDS) performed on Bray-Curtis dissimilarities allowed to analyse the beta-diversity and potential changes in the communities' composition over time. Indeed the alpha diversity indices might not reflect changes in diversity in the cases

<figure><img src="image_8.png" /><figcaption>**Fig. 3.** Evolution of soil functions linked to microbial biomass (A) and the nitrogen cycle (B) at different time points after restoration. The top panels are radar plots, normalised by the value at T0 (dark brown).</figcaption></figure>
where some species are replaced by others, leading to the same richness, but a difference in community structure. When superimposing the NMDS with environmental and soil function data, microbial diversity appears mostly driven by soil water holding capacity (WHC), climate parameters (rainfall), enzymatic activities (urease), and fungi abundance (qPCR 18S). Possibly, fungi diversity would provide additional insight into the impact of restoration on soil microbial communities.

The bacterial community changed significantly in the first year, between T0 and T13, remained comparable for the three following years (i. e., at time points T13, T24 and T36), and finally, another shift in community composition was observed at T60 (Fig. 4).

The first change in beta-diversity from T0 to T13 can be mainly attributed to a loss in Firmicutes, Desulfobacterota and Crenarchaeota, which showed higher relative abundances at T0. The next shift in community composition at T49 observed in the NMDS could be driven by the change in environmental conditions that were more anoxic due to high rainfall as discussed above. Finally, the modification of bacterial communities at T60, might be due to a change in ecological vegetation succession. Indeed, an increase in litter was observed, as well as a reduction in herbaceous plants, with a two-fold decrease in above-ground biomass (AGB) between T49 and T60 (Table S3).

# 4. Discussion

## 4.1. Reduction of particle and contaminant exports after restoration at the watershed scale

Although subject to variability because of spatial differences in slope and extent of vegetation cover, an estimation of particle export from the gold mining flat can be provided. When reported at the scale of the mined area, annual fine particle specific fluxes represent 191 ± 82 t. $km^{-2}.yr^{-1}$ at T0 just after restoration. This is much higher than what was observed in a formerly gold-mined watershed (50 $km^2$) in French Guiana (10–70 t.km$^{-2}.yr^{-1}$ (Hellal et al., 2020)). After two years (T49), the specific particle fluxes were greatly reduced (200 fold, to 1.1 ± 1.2 t. $km^{-2}.yr^{-1}$).

Annual THg specific fluxes showed the same trend in temporal

evolution as particle fluxes and represented 31 ± 15 µgHg.m$^{-2}.yr^{-1}$ at T0, which is more important than what was observed on a small pristine watershed (1.6 $km^2$) in the Brazilian Amazon (3–9 µgHg.m$^{-2}.yr^{-1}$ (Fostier et al., 2000)), but lower than a small (1 $km^2$) formerly gold mined watershed in French Guiana (93 ± 37 µgHg.m$^{-2}.yr^{-1}$ (Guédron et al., 2011)). The reduction in particle fluxes over time results in a reduction of THg fluxes which reached 0.17 ± 0.16 µgHg.m$^{-2}.yr^{-1}$ two years after restoration. This is even lower than the fluxes observed in a small (5 $km^2$) formerly gold-mined watershed in French Guiana during the dry season (0.584 µgHg.m$^{-2}.yr^{-1}$ (Hellal et al., 2020)). Additionally, the mine runoff water appears significantly depleted in THg (p = 0.002) in comparison with the Hg concentration in rainfall samples on the same mining site (3.1 ± 1.2 ng.L$^{-1}$). This indicates that Hg from rainfall might adsorb to soil particles, thereby immobilising it on the mining site, consistently with other sites in French Guiana (Guédron et al., 2011).

The annual Hg influx from rainfall amounts to 8 µgHg.m$^{-2}.yr^{-1}$, hinting towards a net gain of Hg on the mining site, two years after rehabilitation. Hence, the mine soil has evolved from being a source of particulate Hg for downstream rivers during the first year, to becoming a sink able to trap and immobilise Hg, as the particles are stabilised and erosion is reduced. However, in these downslope mining soils depleted in iron oxides and dominated by kaolinite, which is known to be a minor carrier phase for Hg in tropical soils (Guédron et al., 2009; Gurjão et al., 2010), organic matter is the main carrier phase for Hg. Hence, a decrease in carbon concentration over time, as observed in these soils, might also reduce the capacity of the soil to act as a sustainable sink for Hg. Similarly, the reduction of particle exports also induced a reduction in specific yearly Pb and As export from 1500 at T0 down to 17 µg.m$^{-2}.yr^{-1}$ at T49, and from 121 down to 1.2 µg.m$^{-2}.yr^{-1}$ respectively. To the best of our knowledge, no other studies have assessed the export of either of these two contaminants in the Amazon region during mining operations or after the restoration process. The depletion in iron oxides in the downslope mined soils will possibly also affect the retention capacity of the soils for these elements, as both Pb and As are known to show a strong affinity for iron oxides and organic matter (da Silva Junior et al., 2019; Rocha and Mahler, 2023).

The reduction in particle export over time after restoration is mainly

<figure><img src="image_9.png" /><figcaption>**Fig. 4.** NMDS ordination of bacterial diversity profiles from the mining site soils during time after restoration. Significant environmental variables (p < 0.05) are represented by vectors fitted to the NMDS ordination that indicate the direction and magnitude of the change of the variable. (WHC: water holding capacity, URE: urease enzyme activity, 18S: fungi 18S gene qPCR).</figcaption></figure>
<figure><img src="image_10.png" /><figcaption>**Fig. 5.** Synthesis of parameters impacting the restoration of a gold mined soil. Natural parameters are depicted in yellow circles, and human intervention is in red circles. Red arrows represent negative impact and green arrows positive impact.</figcaption></figure>

due to the colonisation of the mining site by herbaceous plants, which provides multiple positive feedback on soil stabilisation (Fig. 5). This induces a stabilisation of soil through the increase in root density, which increases the soils' shear strength (Katuwal et al., 2013), with herbaceous plants being particularly efficient in soil stabilisation due to their great density of very fine roots (Trükmann et al., 2009; Tufekcioglu et al., 1998). In addition, the presence of above-ground plant biomass (AGB) plays a role in rain interception and reducing splash erosion due to the impact of raindrops (Yao et al., 2018), which can be a major component of erosion (Angulo-Martínez et al., 2012). The increase in above-ground biomass is also linked to a reduction in total water exported during a rain event, which can be explained by a reduction of the overland flow velocity, and thus a better water infiltration in the soil. Furthermore, the appearance of a biological soil crust observed on the surface of the soil might have contributed to the prevention of soil erosion, in combination with herbaceous plants and their roots. These "soil biocrusts" are known to help stabilise soils and can play a role in restarting the C and N cycles (Antoninka et al., 2020), and while they are known to develop preferentially in dry climates, biocrusts have been studied in tropical regions as well (Oliveira and Maciel-Silva, 2022; Seitz et al., 2017).

## 4.2. Potential barriers and drivers for the recovery of soil functions

The lack of evolution in the indicators other than arylsulfatase activity could hint towards a lack of sensitivity of these parameters, but more likely the gold mining flat does not undergo a significant change in soil structure and activity over this timespan. The recovery of soil functions is, however, dependent on other parameters such as access to water (climate) and to nutrients (C and N), and the presence of plant-microbial consortia that may regulate the re-establishment of soil functions on the restored mining flat.

Rainfall participates in microbial community changes on the gold mining site (NMDS, Fig. 4), and has an impact on microbial biomass and communities (Ren et al., 2018). The bacterial communities for time point T49, showed an increase in the presence of classes linked to anoxic

environments, like *Desulfuromonadia* (Greene, 2014) and *Anaerolinae* (Yamada and Sekiguchi, 2018), which might be explained by wetter conditions, with the 30 days previous to sampling accumulating 357 mm, corresponding to a 20 % rainfall increase compared to the year before, and 150 % compared to the following year, resulting in soil saturation and potentially anoxia. Indeed, the two first years (until May 2023) of this study were dominated by stronger rainfall (average: 237 ± 135 mm/month) typical of "La Niña" events, whereas the last year (2023–2024) was drier (average: 146 ± 106 mm/month) and typical of an "El Niño" episode, resulting in a drop of 40 % of the previous year's rainfall. Consequently, a significant percentage of planted trees were lost in this last year, and this might also have hampered soil microorganism activity and development.

The decrease of soil C and N content over the studied period can be attributed to the strong development of pioneer herbaceous plants, which have likely consumed the few available resources initially present in the soils. Indeed, of the herbaceous plants identified on the mining site, none were part of the Fabaceae family (Table S4) and are thus not known to be able to fix nitrogen through symbiosis with bacteria. This might lead to a competition for resources in the soil, and this depletion of nitrogen by the herbaceous plants could inhibit the efficient recovery of microbial communities, and thus of soil functions performed by these communities. To assess whether the nitrogen pools are transferred from the soil towards the plant biomass, the total soil N budget in the soil's first 20 cm was compared to the plant (root + above-ground biomass) on that same surface. N concentrations in plants are estimated to be ~1 % (Nardoto et al., 2008; Xiong et al., 2021). Considering the increase of biomass over time, the proportion of N transferred into the biomass through time should not exceed 10–15 % (Fig. S6), pointing towards a stagnant total pool of nitrogen, with rapid recycling of fresh organic matter by the herbaceous plants. It is worth mentioning that only the herbaceous biomass was taken into account in this budget calculation. Hence, on the tree-covered site, a significant proportion of nutrients is stored in the planted trees, meaning that the N budget in plants is underestimated on that site. If this lack of nutrients induces the observed stagnation of microbial activity, it is likely that at a later stage of tree
development, the addition of litter will feed the soil microbial biomass with nutrients, and restore the carbon and nitrogen cycle (as well as other nutrients). It was only at the last time point (T60) that litter began to accumulate under the planted trees, with no visible impact yet on soil nutrients or soil functions. Litter decomposition has been shown to be slower in young secondary forests than in Amazonian primary forests, but with around 30 % loss of litter biomass occurring in the first month after litter-fall (da Silva et al., 2018). However, the quantity and diversity of microorganisms and soil fauna are known to affect the rate of litter decomposition (Krishna and Mohan, 2017), so a lack of the appropriate microorganisms might also explain the absence of litter-fall impact on soil functions.

Another possible explanation for the depletion of C and N in soils is the loss through downward or lateral leaching. Indeed, landscaping during the restoration phase of a gold mine does not restore the soils' initial structure, and results in the creation of a sometimes heterogeneous anthroposol, which no longer benefits from its initial hydric properties regulating water infiltration and retention. As the alluvial mining sites are located in lowland areas, they are prone to undergo major water table oscillations, which can result in the downward or lateral translocation of nutrients, and also possibly the dissolution of their carrier phases (clay minerals), which would otherwise be able to form clay-humic complexes (Fritsch et al., 2011). The loss of these carrier phases might then lead to the loss of the elements/nutrients that would be essential for tree growth and microbial communities.

Leguminous trees of the Fabaceae family, known to form nodules for symbiosis with nitrogen-fixing bacteria (diazotrophs), have been planted during the restoration of the gold mining flat. However, this does not result in a significant input of nitrogen in the soils as discussed above. This might be due to the fact that all the nitrogen is rapidly up-taken into the biomass, but another explanation can be the sheer absence of diazotrophs in these soils. To estimate the presence of N-fixing diazotrophs in the restored mined-soil, the microorganisms detected in the soils were compared to a database of organisms containing the gene responsible for N-fixation (*nifH*) on KEGG. Databases were only compared at the genus level because the sequences in the restored mined-soils were not determined until the species level. This might lead to an overestimation of the presence of N-fixing organisms, as the presence of a N-fixing species does not mean that all the species in the genus are capable of N-fixation. However, this overestimation should be similar between all samples, allowing us to follow a potential evolution over time. An increase of diazotroph presence was observed at T49 and T60 (Fig. S9), mostly due to *Paludibacter* and *Methylocystis*. Even when those genii are removed, the increase is still visible, albeit less substantial. The mining site restoration thus seems to be accompanied by a return of N-fixing species, especially as the planted trees are from the Fabaceae family, and are able to form root nodules and symbiotic relationships with diazotrophs. In contrast, a lower percentage of N-fixing bacteria is observed in the forest soils, which could be explained by more abundant N in these soils, and by other sources of available nitrogen (e.g., decomposed OM). As a result, N-fixation does not give a competitive advantage compared to other genii in a functional primary or secondary forest.

The forest edge effect observed on the landscaped site led to a differential revegetation on the three erosion plots, due to an input of nutrients, microorganisms, and seeds from the forest, which also had a visible impact on different soil function indicators (ARS, DNA, genes linked to the nitrogen cycle, Fig. S2). The facilitation of the closure of mines by the forest edge effect has been investigated in more detail in the Amazon and elsewhere (Garate-Quispe et al., 2023; Martínez-Ruiz et al., 2021; Parrotta and Knowles, 2001).

## 4.3. Recommendations and perspectives for goldmining sites restoration

Using the data presented here, a timeline for the restoration of an alluvial gold mining site can be proposed. Indeed, after landscaping and replanting, rapid soil colonisation by herbaceous plants and soil biocrust,

occurs within the first year and leads to a sharp decline in particle and contaminant export. The planted trees then grow to 3–5-meters high leading to a tree-cover succession stage within 3 years. However, even 5 years after restoration, no clear recovery of soil functions was observed. This can be attributed to the depletion of soil nutrients by the herbaceous plant growth and to nutrient leaching which hinders the completion of the next succession stage where a secondary forest would appear. In order to promote an efficient rehabilitation and closure of the mining site, the importance of rapid landscaping after the exploitation phase is paramount. Indeed, if there is little time between the exploitation of the mining site and its restoration, the organic layer stored by the miners will be less susceptible to leaching and loss of nutrients, microorganisms, and seeds, all of which are essential for the adequate closure of a mining site. Hence, rehabilitation integrated into the progress of extraction activities seems to be the best strategy. A fast and efficient recolonisation by herbaceous plants leading to reduced particle and contaminant and more specifically Hg export is in line with the Minamata convention on mercury, which states that national action plans (NAPs) should include “Strategies for promoting the reduction of emissions and releases of, and exposure to mercury in artisanal and small-scale gold mining and processing” (Minamata_Convention_on_Mercury, 2023). Several organisations work across the world and more specifically in the Amazon region to raise awareness in gold-miner communities and to limit the impact of gold-mining on the environment (planetGOLD, 2023).

In addition, many mines in French Guiana use leguminous plants to help restart the nitrogen cycle through symbiotic nitrogen-fixing microorganisms (Couic et al., 2022). These leguminous trees appear to have an impact on the soil microbial communities as it was observed that more N-fixing microorganisms are present on the tree-covered site. However, these microorganisms appear only 4 to 5 years after restoration, which was also coupled with an increase in litterfall and a decrease in herbaceous cover after 5 years. Possibly, at this succession stage, the planted trees provide shade, which reduces the ability of herbaceous plants to grow and also impacts the soil microbial communities. This reflects a start in adaptation to the changing environment of the rehabilitated mining site and gives promising perspectives for the future recovery of soil functions and microbial communities in mining site soils. Additional inoculation of the tree shoots with N-fixing bacteria could improve restoration and the recovery of the N-cycle. Finally, the variability over the whole mining site needs to be better assessed, as differential parameters (physicochemical, slope, edge effect, ...) have the possibility to influence the development of microbial communities creating patches of diversity and “efficient rehabilitation” but only on specific locations on the mining site. Taking advantage of the positive influence of the edge effect could include focusing the rehabilitation effort on punctual niches as “hotspots” for microbial dispersion, or also conducting mine work over narrower widths in order to achieve a larger forest edge over mined area ratio.

The time frame of this study appeared to be too short to detect any significant increase in soil functions, possibly because some of the parameters discussed here are hindering the refunctionalisation of the soils. Further studies thus need to perform monitoring of a mining site over a longer period of time (in the decade time frame), to document the transition towards the restarting of the carbon and nitrogen cycles through the input of organic matter from litter-fall back to the soils. This tipping point is essential for the recovery of a secondary forest and appears to lie in the decades’ timespan. On an even larger timescale, climate change is believed to lead to lower rainfall in tropical regions (Neelin et al., 2006), which might add to the difficulty of restoring efficiently a mining site.

The application of mining site restoration to illegal gold mines has to be studied, as in the case of illegal gold mining, Hg is used for Au amalgamation and leads to contamination of the surrounding water and the soils. It is thus not trivial to decide whether an illegal mining site should be restored and how, as restoration might lead to the remobilisation of the anthropogenic Hg in soils during the landscaping process.
## CRediT authorship contribution statement

**Naomi Nitschke:** Writing – review & editing, Writing – original draft, Visualization, Validation, Investigation, Formal analysis, Data curation, Conceptualization.
**Jennifer Harris:** Writing – review & editing, Writing – original draft, Visualization, Validation, Supervision, Resources, Project administration, Methodology, Investigation, Funding acquisition, Data curation, Conceptualization.
**David Amouroux:** Writing – review & editing, Validation, Supervision, Resources, Project administration, Methodology, Investigation, Funding acquisition, Data curation, Conceptualization.
**Emmanuel Tessier:** Validation, Supervision, Methodology, Formal analysis, Data curation.
**Cédric Legout:** Writing – review & editing, Validation, Methodology, Investigation, Data curation.
**Sylvain Campillo:** Writing – review & editing, Validation, Supervision, Methodology, Formal analysis.
**Stéphane Taravella:** Resources, Funding acquisition.
**Stéphane Guedron:** Writing – review & editing, Writing – original draft, Visualization, Validation, Supervision, Resources, Project administration, Methodology, Investigation, Funding acquisition, Data curation, Conceptualization.

## Funding sources

This work is a contribution to the project "ReHab" funded by the ANR [PI, S. Guédron is part of Labex OSUG (ANR10 LABX56)], and the project "Instrumentation d'un hydrosystème en contexte d'orpaille pour évaluer les effets de la réhabilitation minière sur les flux de MES et de contaminants (Hg, Pb, Cd et As) au cours de la réhabilitation" funded by the DGTM (Direction Générale des Territoires et de la Mer de Guyane, service prévention des risques et industries extractives) [2022/2024]. This work also was supported by the ARNT-CIFRE [2021/0723], by the CTG (Collectivité Territoriale de Guyane) [project name: ODYSSEE], and by the company SAS GAIA. Carbon analyses were performed as a part of the ANR-22-PEXF-006 - CarboNium project of the FairCarBoN exploratory PEPR and has received State aid managed by the Agence Nationale de la Recherche under the France 2030 program, with the reference ANR.

## Declaration of competing interest

The authors declare the following financial interests/personal relationships which may be considered as potential competing interests: Naomi Nitschke reports financial support and travel were provided by GAIA-SAS. Naomi Nitschke reports a relationship with GAIA-SAS that includes: employment. If there are other authors, they declare that they have no known competing financial interests or personal relationships that could have appeared to influence the work reported in this paper.

## Acknowledgments

We thank Océane Asensio (UPPA-Tech, IPREM UPPA/CNRS), Cindy Arnoldi (LECA, UGA), Catherine Joulian (BRGM) and Hafida Tris (BRGM) for analytical support in the laboratory. Thanks to Mélanie Roy (IRD) for help in plant identification. Finally, we thank Akram Ali (SAS GAIA) for field assistance.

## Appendix A. Supplementary data

Supplementary data to this article can be found online at https://doi.org/10.1016/j.scitotenv.2025.180150.

## Data availability

Data will be made available on request.

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