Dr Denis Habauzit
email: denis.habauzit@univ-rennes1.fr

Catherine Martin

Gwenneg Kerdivel

Farzad Pakdel
email: farzad.pakdel@univ-rennes1.fr

Rapid assessment of estrogenic compounds by CXCL-test illustrated by the screening of the UV-filter derivative Benzophenones

Keywords: Estrogens, detection, CXCL-test, endocrine disrupting chemicals, benzophenone, xeno-estrogens

Rapid assessment of estrogenic compounds by CXCL-test illustrated by the screening of the UV-filter

derivative benzophenones.

Introduction

Human activities, due to industrialization and to drugs consumption, have led to the increase of diverse chemical releases in environment, especially in surface water. This increase is partially due to the incomplete elimination of the contaminants in waste water treatment plan. Some of these chemicals evidenced unpredicted endocrine disrupting activities. The main difficulty in this research area is that the disruption could occur in large biological tissues, functions, pathways or proteins (Kerdivel et al., 2013a;[START_REF] Yoon | Estrogenic endocrine-disrupting chemicals: molecular mechanisms of actions on putative human diseases[END_REF]. Therefore the assessment of endocrine disrupting activity for many compounds was always performed a posteriori either after several years of the use of compounds or because of an evidence of a health effect, such as the use of Diethylstilbestrol or Bisphenol A. Moreover, the assessment is sometimes performed after the detection of a high concentration of several compounds in environment linked with adverse effects in wildlife evidenced through morpho-anatomical modifications or expression of feminine protein marker in male fish. Since, many comity or regulations have aimed to develop early evaluation of the new chemical such as EDSTAC (now EDSP) and Reach program. Several methods have been established over the last 3 decades in order to assess endocrine disrupting chemical (EDC) activities, especially estrogenic activity [START_REF] Leusch | Tools to Detect Estrogenic Activity in Environmental Waters[END_REF][START_REF] Scholz | Alternatives to in vivo tests to detect endocrine disrupting chemicals (EDCs) in fish and amphibians--screening for estrogen, androgen and thyroid hormone disruption[END_REF]. These methods specifically target several critical points that have been identified in fundamental studies in the endocrine function such as development of uterus [START_REF] Isenhower | Absence of estrogenic activity in some drugs commonly used during pregnancy[END_REF], estrogen-induced proliferation of cell lines [START_REF] Soto | The E-SCREEN assay as a tool to identify estrogens: an update on estrogenic environmental pollutants[END_REF], expression of estrogen-regulated genes (vitellogenin) [START_REF] Flouriot | Influence of xenobiotics on rainbow trout liver estrogen receptor and vitellogenin gene expression[END_REF] or molecular interaction with transcription factors [START_REF] Habauzit | SPR-based biosensors: a tool for biodetection of hormonal compounds[END_REF]. But the evaluation of the estrogenic activity of a molecule raises several questions about the relevance of the target, the considered material (organ, tissue, cells or molecules), and the analysis time and cost.

CXCL12 is a cytokine coupled with two G protein coupled receptors (CXCR4 and CXCR7) that is implicated in the proliferation and migration of breast cancer cells [START_REF] Boudot | Differential estrogen-regulation of CXCL12 chemokine receptors, CXCR4 and CXCR7, contributes to the growth effect of estrogens in breast cancer cells[END_REF]. E2-induced CXCL12 expression and its impact on cell proliferation has then been illustrated by our team and others [START_REF] Boudot | Differential estrogen-regulation of CXCL12 chemokine receptors, CXCR4 and CXCR7, contributes to the growth effect of estrogens in breast cancer cells[END_REF][START_REF] Glace | Estrogeninduced stromal cell-derived factor-1 (SDF-1/Cxcl12) expression is repressed by progesterone and by Selective Estrogen Receptor Modulators via estrogen receptor alpha in rat uterine cells and tissues[END_REF][START_REF] Hall | Stromal cell-derived factor 1, a novel target of estrogen receptor action, mediates the mitogenic effects of estradiol in ovarian and breast cancer cells[END_REF][START_REF] Pattarozzi | 17beta-estradiol promotes breast cancer cell proliferation-inducing stromal cell-derived factor-1-mediated epidermal growth factor receptor transactivation: reversal by gefitinib pretreatment[END_REF]. Moreover, we have demonstrated that CXCL12 secretion is an early indicator of estrogen-dependent proliferation of breast cancer cells (MCF7 and T47D). The knowledge of estrogen-dependent regulation of this cytokine led to the development of a fast test for the assessment of estrogenic effect of molecules [START_REF] Habauzit | Development and validation of a test for environmental estrogens: Checking xeno-estrogen activity by CXCL12 secretion in BREAST CANCER CELL LINES (CXCL-test)[END_REF]. By using CXCL12 secretion for EDC-A c c e p t e d m a n u s c r i p t evaluation, Hall and Korach demonstrated that some EDCs also induce growth of ovarian cancer cells through the estrogen receptor (ER)-dependent induction of CXCL12 expression [START_REF] Hall | Endocrine disrupting chemicals promote the growth of ovarian cancer cells via the ER-CXCL12-CXCR4 signaling axis[END_REF]. CXCL-test is faster (14 hours) than the classical proliferation test (5-7 days) and sensitive enough for the assessment of both strong and weak estrogenic compounds.

Even if a lot of methods for the evaluation and the quantification of EDCs exist, few tests are using natural endogenous genes expression with a low detection limit for the evaluation of the proliferative ability of molecules. Moreover, due to the diversity of estrogenic compounds' actions and pathways, numerous methods have to be developed for the entire evaluation of the molecule action. For that reason, we aimed to ameliorate and optimize the CXCL-test by scaling it up to 96 wells plate and by reducing intermediate steps. To validate these optimizations and illustrate their relevance for risk assessment, we assessed a new EDC family with estrogen-like activity, the benzophenones. They are used in several cosmetic products and they possess ultraviolet (UV) filter activities. Because of their lipophilic properties, they penetrate easily in bodies and can be found in human fluid such as urine [START_REF] Vela-Soria | A multiclass method for the analysis of endocrine disrupting chemicals in human urine samples. Sample treatment by dispersive liquid-liquid microextraction[END_REF], women milk [START_REF] Rodríguez-Gómez | A multiresidue method for the determination of selected endocrine disrupting chemicals in human breast milk based on a simple extraction procedure[END_REF]. The diversity of benzophenone's use makes that large amount of these compounds are found in the waste water treatment plant and constitute a risk for wildlife [START_REF] Jurado | Urban groundwater contamination by residues of UV filters[END_REF]. As the treatment plant eliminates only partially these compounds, they are found to high concentration (up to mg/L) in rivers all over the world [START_REF] Tsui | Occurrence, distribution and ecological risk assessment of multiple classes of UV filters in surface waters from different countries[END_REF] and also in sediment [START_REF] Gago-Ferrero | Fast pressurized liquid extraction with in-cell purification and analysis by liquid chromatography tandem mass spectrometry for the determination of UV filters and their degradation products in sediments[END_REF]. There are increasing concerns on the potential endocrine disrupting effect of these molecules. In this study, we show that among the 10 compounds evaluated, all have an estrogenic activity.

A c c e p t e d m a n u s c r i p t

Material and methods

Materials

Actinomycin D, benzophenone (BP), bovin serum albumin (BSA), 17--estradiol (E2), 17--ethynylestradiol (EE2), 2,4-dihydroxy-benzophenone (BP1), genistein (Gen), 2,2',4,4'tetrahydroxy-benzophenone (BP2), 2-hydroxy-4-methoxy-benzophenone (BP3), 4-hydroxybenzophenone (4BP), 4,4'-dihydroxybenzophenone (44'BP), 2,3,4-trihydroxy-benzophenone (234BP), 2,3,4,4'-tetrahydroxy-benzophenone (2344'BP), 2,2'-hydroxy-4-methoxy-benzophenone 6 (BP8), 2,4,4'-trihydroxy-benzophenone (THB), progesterone (P4) and testosterone (T) were acquired from Sigma-Aldrich Co.

Cell Culture and treatments

T47D cells, obtained from ATCC, were routinely maintained in RPMI (Invitrogen, Life technologies, Saint Aubin, France) supplemented with 10% fetal bovine serum, FBS (Biowest, Nuaillé, France) and antibiotics containing penicillin, streptomycin and Amphotericin (Invitrogen, Life technologies, Saint Aubin, France) at 37°C in 5% CO2. Before treatment cells were plated in DMEM phenol red-free (Gibco, Life technologies, Saint Aubin, France) containing 5% dextran treated charcoal-stripped FBS (dsFBS) provided by Biowest (Biowest, Nuaillé, France). For treatment, 0.1 % ethanol (control) and each compound (E2, EE2, Gen, BP, BP1, BP2, BP3, 4BP, 44'BP, 234BP, 2344'BP, BP8, THB, P4 and T) was diluted with DMEM phenol red-free containing 5% dsFBS at the appropriate concentration.

ELISA assay

Cells were cultured in 96-well plates containing either 4.10 4 cells/well in DMEM phenol red-free medium containing 5% of dsFBS. After 24 hours, medium was replaced and cells were then treated with DMEM phenol red free medium with 5% of dsFBS containing the tested molecules during 14 hours. Then, 60 µl of supernatant were equally diluted with PBS containing 1% BSA. The quantification of CXCL12 secretion was then performed by ELISA (Quantitine kit; R&D Systems, Mineapolis, MN, USA).

A c c e p t e d m a n u s c r i p t

Then in order to decrease the number of cells used for one experiment and to avoid the intermediate dilution step, several cell density per well were tested. Between 2500 to 40000 cells per well were evaluated.

Finally Ten thousands cells per well, cultured in 96-well plate, were chosen for the evaluation of BP estrogenic potency. The medium was then replaced with 100 µl of DMEM phenol red-free medium containing 5% of dsFBS during 24 hours. Then 100 µl of DMEM phenol red free containing treatment were added. After 14 hours of incubation, 100 µl of the supernatant was directly analyzed by ELISA. For each experiment, a standard curve was generated. ELISA was revealed with the CXCL12 Duoset Elisa kit from R&D Systems. The absorbance at 450 and 570 nm was obtained by using a 96-well iMARK Absorbance microplate reader (Biorad, Life Science, Marnes-la-coquette, France).

RT-PCR assays

T47D cells were seeded in 6 well-plate in a density at 1.10 5 cells/well in DMEM phenol-red free medium containing 5 % of dsFBS. After 24 hours, cells were treated with 10 -6 M of each BP and 10 -8 M of E2 and 0.1% ethanol as control during 24 hours. For each treatment, total RNA was extracted with Trizol reagent (Invitrogen) according to supplier's instructions. RNAs were then reverse transcribed in cDNA by MMLV (Promega, Charbonnère, France) using random hexamers (Promega). Quantitative RT-PCR were performed on BioRad MyiQ with iQ Sybr Green Supermix (BioRad, Hercules, CA, USA) using primers for CXCL12, Rev: GCCTCCATGGCATACATAGG, Fwd: CTCCTGGGGATGTGTAATGG and for GAPDH as control, Rev: GGGCATCCTGGGCTACACTG, Fwd: GAGGTCCACCACCCTGTTGC.

Statistics

Statistical analyses were performed using Mann and Whitney test on Biostatgv online tool (http://marne.u707.jussieu.fr/biostatgv/?module=tests). Differences were considered as significant if p value < 0.05.

A c c e p t e d m a n u s c r i p t

RESULT

Effect of classical estrogenic compounds on the secretion of CXCL12 in 96 well-plate

In previously published method, 24 well-plates were used. By scaling up this method to 96 well-plates, we aimed to screen more conditions during the same experiment. We therefore evaluate the effect of classical estrogenic compounds such as the strong agonist 17-α-ethynyl estradiol (EE2), the natural hormone 17-βestradiol (E2) and the weak agonist genistein (Gen). Each compound induced an increase of the secretion of CXCL12 when compared to the control ethanol (Fig. 1). E2 and EE2 induced up to 5 to 6-fold increase of CXCL12 secretion. A decrease of the CXCL12 secretion was nevertheless observed for concentration of EE2 up to 10 -8 M as previously published [START_REF] Habauzit | Development and validation of a test for environmental estrogens: Checking xeno-estrogen activity by CXCL12 secretion in BREAST CANCER CELL LINES (CXCL-test)[END_REF]. Gen also increased the CXCL12 secretion.

Consistently with its well described weak agonistic potency, Gen only induced effect at doses higher than 10 -8 M. The EC50 of each curves was determined and was respectively of 3.15×10 -13 M for E2, 2.14×10 -14 M for EE2 and 7.17×10 -8 M for Gen (Table 1). The first effects of molecules on CXCL12 secretion were observed at 10 -14 M E2 (3 fold increase) and 5×10 -16 M of EE2 (2-fold increase). A supernatant dilution step was previously performed to keep the ELISA signal in the linear range of the calibration curve. In this second step of the method optimization, the intermediary dilution of the supernatant in PBS 1% BSA was removed by decreasing cell number. Different cell density from 2500 cells/well to 40000 cell/well (Fig. 2) were cultured on 96-well plates. Cells were exposed to 10 -8 M E2 or ethanol, as control. The secretion of CXCL12 in 100 µl of supernatant was monitored and represented as a fold change between exposed to ethanol and E2 (Fig. 2). The 2×10 4 and 4×10 4 cell numbers permitted to visualize the E2 effect with fold change of 2.4 compared to the ethanol control. The presence of an equivalent fold change in these two cellular densities raises the question of the signal saturation, even if for 2×10 4 cells per well, the ELISA is performed in the linear range of the calibration curve. Therefore 1×10 4 cells per well were preferred. With a fold change of 1.9, this cell number allowed significantly to evaluate the estrogenic potency of compounds without saturation signal risk.

Specificity of the estrogenic sensitivity of the CXCL12 secretion

Two well-known endocrine compounds with non-estrogenic activity were tested on the ability to induce the CXCL12 secretion. These two compounds were the male hormone testosterone (T) and the women menstrual and pregnancy hormone progesterone (P4). The stimulation by a dose effect of these two compounds did not A c c e p t e d m a n u s c r i p t modify the secretion of the CXCL12 (Fig. 3). Altogether these results evidenced that the response is specific to estrogenic stimulation.

Fig. 3: Assessment of the estrogenic properties of dose effect of A) testosterone (T) and B) progesterone (P4) by CXCL-test

Use of the optimized CXCL-test for the assessment of estrogenic potency of Benzophenones

According to our developed ELISA method, all 10 tested benzophenones exhibit estrogenic properties in a dose-dependent manner (Fig. 4). Nevertheless, among these 10 molecules, two groups of estrogenic compounds can be delineated. The first one contains 6 compounds (BP, BP1, BP2, BP3, 234BP and 2344'BP)

A c c e p t e d m a n u s c r i p t that have a weak estrogenic activity with EC50 near 10 -5 M (table 1). The only active dose is 10 -5 M and the maximum response is lesser than 70% of E2 effect. The 4 last compounds (4BP, 44'BP, BP8 and THB) have medium estrogenic properties with active concentration from 10 -6 M and the maximum response is between 80% and 110% of E2 effect. The most powerful estrogenic compounds are 44'BP (near 100% of E2 effect) and BP8 (110% of E2 effect).

Compounds EC50 (M) E2

3.15×10 A c c e p t e d m a n u s c r i p t

Validation of CXCL-test by quantitative PCR

To confirm that stimulation of CXCL12 secretion by BPs is only due to the stimulation of gene expression and not to a transient increase of CXCL12 protein release by action on exocytose, RT-qPCR method was also used. After 24 hours of 10 -6 M BP treatments, T47D RNA were extracted and then analyzed by qPCR.

As shown in Fig. 5A, CXCL12 mRNA obtained after 24 hours of treatment correlated very well with CXCL12 secretion obtained after 14 hours of treatment. The Pearson correlation is 84.15% with a p < 0.001. Of note is that at 10 -6 M, BP molecule did not show estrogenic activity in both assays while 4BP and BP8 were among the most powerful estrogenic compounds. The differences observed between qPCR and ELISA approaches are likely due to the treatment time as well as the sensitivity of the method.

To further investigate whether or not the effect of BPs on CXCL12 release is directly dependent on CXCL12 gene transcriptional activity, T47D cells were treated with different doses of BPs, in presence of an inhibitor of the transcription, the actinomycine D. As shown in Fig. 5B, Actinomycin D inhibited the BPdependent CXCL12 secretion for both E2 and BP molecules at their effective doses (10 -6 M or 10 -5 M). No transcriptional independent release of CXCL12 was observed and no significant differences between control ethanol and BP treatments were observed. The BP-dependent CXCL12 secretion is therefore due to the estrogenic effect of molecules and not of an increase in CXCL12 release in cell supernatant. A c c e p t e d m a n u s c r i p t

DISCUSSION

Since decades several methods have been developed for the evaluation and the quantification of endocrine disrupting chemicals, using their ability to induce proliferation, expression of reporter genes or endogenous genes or any other characteristics of these compounds (Kerdivel et al., 2013a;[START_REF] Leusch | Tools to Detect Estrogenic Activity in Environmental Waters[END_REF][START_REF] Scholz | Alternatives to in vivo tests to detect endocrine disrupting chemicals (EDCs) in fish and amphibians--screening for estrogen, androgen and thyroid hormone disruption[END_REF]. These methods used (i) whole animal with uterotrophic test in mouse [START_REF] Isenhower | Absence of estrogenic activity in some drugs commonly used during pregnancy[END_REF] or vitellogenin test in fish [START_REF] Flouriot | Influence of xenobiotics on rainbow trout liver estrogen receptor and vitellogenin gene expression[END_REF], (ii) cell-based in vitro assays such as proliferation test [START_REF] Soto | Control of cell proliferation: evidence for negative control on estrogen-sensitive T47D human breast cancer cells[END_REF], differentiation test [START_REF] Habauzit | Differentiation of PC12 cells expressing estrogen receptor alpha: a new bioassay for endocrine-disrupting chemicals evaluation[END_REF] or reporter assays and (iii) molecular in vitro assay with receptor binding assay [START_REF] Habauzit | Determination of estrogen presence in water by SPR using estrogen receptor dimerization[END_REF]. Each of these methods has their own advantages such as integrated response, physiological response, specific response or fast response. Here we have aimed to optimize the CXCL-test in order to increase the number of molecules to be tested in a single assay. We also evidenced the low detection limit of the CXCL-test and that the use of natural endogenous gene expression could be an early indicator of the proliferation capacity of a molecule. This development opens the possibility to use this test as a high throughput assay for estrogenic compounds screening. This method is fast and alternative to animal test in agreement with REACH program (Lombardo et al., 2014a(Lombardo et al., , 2014b)). At the moment, one of the main methods used for evaluation of xeno-estrogens properties of chemicals is the ER-positive cell line proliferation test based on Soto et al. work which was originally developed on T47D cell line [START_REF] Soto | Control of cell proliferation: evidence for negative control on estrogen-sensitive T47D human breast cancer cells[END_REF]. The CXCL-test is a method that allows fast assessment of the estrogen-like action of endocrine disruptors. This test constitutes an alternative for proliferative test (14 hours versus 5 days) of ER-positive cell lines (such as MCF7 or T47D) [START_REF] Habauzit | Development and validation of a test for environmental estrogens: Checking xeno-estrogen activity by CXCL12 secretion in BREAST CANCER CELL LINES (CXCL-test)[END_REF]. This test relied on the ability of estrogens to induce the expression and the secretion of CXCL12 in ER-positive cells which has been demonstrated in many tissues and cancers [START_REF] Boudot | COUP-TFI modifies CXCL12 and CXCR4 expression by activating EGF signaling and stimulates breast cancer cell migration[END_REF][START_REF] Roberti | Protein expression changes during human triple negative breast cancer cell line progression to lymph node metastasis in a xenografted model in nude mice[END_REF]. CXCL12 and its two G-coupled receptors CXCR4 and CXCR7 are involved in the proliferation of ER-positive cancer cells [START_REF] Boudot | Differential estrogen-regulation of CXCL12 chemokine receptors, CXCR4 and CXCR7, contributes to the growth effect of estrogens in breast cancer cells[END_REF], and is also a marker of the breast cancer progression [START_REF] Boudot | COUP-TFI modifies CXCL12 and CXCR4 expression by activating EGF signaling and stimulates breast cancer cell migration[END_REF]. Indeed, when breast cancer cell progress they lose the estrogenic sensitivity, the decrease of the local CXCL12 secretion induces EMT and metastasis [START_REF] Boudot | COUP-TFI modifies CXCL12 and CXCR4 expression by activating EGF signaling and stimulates breast cancer cell migration[END_REF][START_REF] Roberti | Protein expression changes during human triple negative breast cancer cell line progression to lymph node metastasis in a xenografted model in nude mice[END_REF]. In this article, we proposed key parameters that should be considered for test improvement. First of all, we intended to increase the number of tested molecules at the same time by adapting the method to the 96 well plates. By screening the classical estrogenic compounds E2, EE2 and Gen, the reproducibility of the method was checked with our previously data [START_REF] Habauzit | Development and validation of a test for environmental estrogens: Checking xeno-estrogen activity by CXCL12 secretion in BREAST CANCER CELL LINES (CXCL-test)[END_REF]. Secondly, in order to avoid the intermediate dilution of the supernatant, the number of cells was decreased to 10000 cells per well. The final method greatly decreases A c c e p t e d m a n u s c r i p t the time and steps needed for the global evaluation of the estrogenic potency of a large number of compounds.

Finally, when compared to the classically used proliferation test, our optimized CXCL-test allows an easier interpretation of the results and the saving of time is really significant even if this test could exhibit a slightly more expensive cost. These kinds of fast and accurate detection methods are clearly important for risk assessment, considering the constant increase of new chemicals released in the environment.

Altogether, these optimized conditions were used for the screening of the benzophenone derivatives that are emerging endocrine disruptors. These molecules are present in some cosmetics and are found in many surface waters around the world with an environmental concentration from ng/L to µg/L [START_REF] Jurado | Urban groundwater contamination by residues of UV filters[END_REF][START_REF] Loraine | Seasonal variations in concentrations of pharmaceuticals and personal care products in drinking water and reclaimed wastewater in southern California[END_REF][START_REF] Molins-Delgado | Ecological risk assessment associated to the removal of endocrine-disrupting parabens and benzophenone-4 in wastewater treatment[END_REF]. These compounds began to be increasingly studied because of their potential risk for human and wildlife (Kim and Choi, 2014).

All tested BPs exhibit estrogen-like activity with various potencies. Indeed, BP, BP1, BP2, BP3, 234BP and 2344'BP showed a weak estrogenic potency even at 10 -5 M. Surprisingly, the BP molecule showed a weak but detectable estrogenic activity in inducing CXCL12 secretion. However our previous docking analysis suggested no direct interaction with ER ligand-binding domain (Kerdivel et al., 2013b). This low estrogenic activity of BP molecule observed at 10 -5 M is probably due to its conversion in 4BP in cell culture [START_REF] Hayashi | Formation of estrogenic products from benzophenone after exposure to sunlight[END_REF][START_REF] Nakagawa | Metabolism and toxicity of benzophenone in isolated rat hepatocytes and estrogenic activity of its metabolites in MCF-7 cells[END_REF]. On the other hand, 4BP, 44'BP, BP8 and THB showed much more potent estrogenic activity inducing a response from 10 -6 M. Among these compounds, 4BP and BP8 were also described as the most potent estrogenic compounds, in regards to previously published results on estrogensensitive reporter assays and on the expression of the CXCL12 gene in MCF7 cell lines (Kerdivel et al., 2013b).

The difference in the estrogenic sensitivity is probably due to the ER/ER ratio which is different in MCF7 and T47D cells. Our previously published data have also evidenced that T47D cells are more sensitive to estrogenic stimulation than MCF7 especially for genistein [START_REF] Habauzit | Development and validation of a test for environmental estrogens: Checking xeno-estrogen activity by CXCL12 secretion in BREAST CANCER CELL LINES (CXCL-test)[END_REF]. Moreover, ER stably expressed in Hela cells (HELN ER assay) exhibited much more sensitivity than ER for several benzophenones such as BP1, BP2 and THB [START_REF] Molina-Molina | Profiling of benzophenone derivatives using fish and human estrogen receptorspecific in vitro bioassays[END_REF].

Under benzophenone treatments, stimulations of the CXCL12 mRNA expression and protein secretion were nearly identical with a correlation of more than 80%. This was also demonstrated by the actinomycin D treatments, showing that the increase of CXCL12 secretion is only due to the increase of CXCL12 expression and not to an increase of CXCL12 release in the supernatant by BP-induced exocytosis. However, because of the A c c e p t e d m a n u s c r i p t duration of treatment (14 hours for secretion and 24 hours for the qPCR) and technique sensitivity, some differences in the amplitude of response were observed. Here we evaluated only the short-term effect of these molecules. However, the benzophenone endocrine activity is currently under question especially about the potential transgenerational effects. For instance, a recent study showed that BP3 has adverse effects in Japanese medaka and some transgenerational disrupting effects in hormonal balance and reproductive function may be noticed when fish were exposure at the level of μg / L to BP3 (Kim et al., 2014).

CONCLUSION

Today thousand chemicals are currently used in common human life without knowing their endocrine disrupting status. Among the number of methods that have been developed last decades, the CXCL-test promotes a very low detection limit for well-known estrogenic compounds E2 and EE2, and used a natural and untransformed gene that is involved in the proliferation and the migration of the breast cancer cells. In this study, we described the optimization of CXCL-test that permits fast and reliable evaluation of the estrogenic potencies of environmental chemicals. We used this optimized test to evaluate the EDC activity of some benzophenone derivatives. These compounds present in environment in high concentration have an estrogenic activity and constitute a risk for human and wildlife.

Fig. 1 :

 1 Fig. 1: CXCL12 secretion under classical estrogenic compounds stimulation in 96 well plate.

Fig. 2 :

 2 Fig. 2: CXCL12 secretion depending upon the cell number coated in the 96 well plate and without PBS/BSA intermediary dilution. (*P<0.05, **P<0.01, ***P<0.005)

Fig. 4 :

 4 Fig. 4: Assessment of estrogenic properties of different doses of benzophenones (BP, BP1, BP2, BP3, 4BP, 44'BP, 234BP, 2344'BP, BP8 and THB) using CXCL-test.

Fig. 5 :

 5 Fig. 5: A) Evolution of CXCL12 secretion and CXCL12 expression measured respectively by ELISA and qPCR after stimulation with E2 at 10 -8 M and benzophenones at 10 -6 M. B) CXCL-test performed on cells treated or not with control ethanol, E2 and effective doses of benzophenones with or without actinomycin D.

Table 1 :

 1 EC50 of each compound used in the evaluation.

		-13
	EE2	2.14×10 -14
	Gen	7.18×10 -8
	BP	1.09×10 -5
	BP1	9.46×10 -6
	BP2	8.29×10 -6
	BP3	9.93×10 -6
	4BP	9.45×10 -7
	44'BP	6.7×10 -7
	234BP	7×10 -6
	2344'BP	8.42×10 -6
	BP8	1.5×10 -6
	THB	5.24×10 -7

Acknowledgements

Authors are grateful to Dr. Frederic Leusch for its help and advices during the evaluation of this article. This work was supported by funds from the French Ministry of Ecology, Energy and Sustainable Development (PNRPE). We also thank CNRS, INSERM and the University of Rennes 1 for supporting our research.

Conflict of interest

The authors declare that they have no conflict of interest.

BIBLIOGRAPHY

A c c e p t e d m a n u s c r i p t


Cristina Botías
email: cristinabotias@gmail.com

Arthur David

Elizabeth M Hill

Dave Goulson

Quantifying exposure of wild bumblebees to mixtures of agrochemicals in agricultural and urban landscapes

Keywords: neonicotinoids, fungicides, pesticide mixtures, wild bumblebees, pesticide exposure

The increased use of pesticides has caused concern over the possible direct association of exposure to combinations of these compounds with bee health problems. There is growing proof that bees are regularly exposed to mixtures of agrochemicals, but most research has been focused on managed bees living in farmland, whereas little is known about exposure of wild bees, both in farmland and urban habitats. To determine exposure of wild bumblebees to pesticides in agricultural and urban environments through the season, specimens of five different species were collected from farms and ornamental urban gardens in three sampling periods. Five neonicotinoid insecticides, thirteen fungicides and a pesticide synergist were analysed in each of the specimens collected. In total, 61% of the 150 individuals tested had detectable levels of at least one of the compounds, with boscalid being the most frequently detected (35%), followed by tebuconazole (27%), spiroxamine (19%), carbendazim (11%), epoxiconazole (8%), imidacloprid (7%), metconazole (7%) and thiamethoxam (6%).

Quantifiable concentrations ranged from 0.17 to 54.4 ng/g (bee body weight) for individual pesticides. From all the bees where pesticides were detected, the majority (71%) had more than one compound, with a maximum of seven pesticides detected in one specimen.

Concentrations and detection frequencies were higher in bees collected from farmland compared to urban sites, and pesticide concentrations decreased through the season. Overall, our results show that wild bumblebees are exposed to multiple pesticides when foraging in agricultural and urban landscapes. Such mixtures are detected in bee tissues not just during the crop flowering period, but also later in the season. Therefore, contact with these combinations of active compounds might be more prolonged in time and widespread in the environment than previously assumed. These findings may help to direct future research and pesticide regulation strategies to promote the conservation of wild bee populations.

Introduction

Bees are exposed to environmental pollutants via contaminated food resources such as pollen, nectar or water [START_REF] Bonmatin | Environmental fate and exposure; neonicotinoids and fipronil[END_REF], and through external contact with aerosols during spraying and contaminated dust emitted during the sowing of dressed seeds as their hairy bodies trap particulate residues [START_REF] Greig-Smith | Incidents of poisoning of honeybees (Apis mellifera) by agricultural pesticides in Great Britain 1981-1991[END_REF][START_REF] Pistorius | Application of predefined doses of neonicotinoid containing dusts in field trials and acute effects on honey bees[END_REF]. Many studies have used honeybees as relevant organisms to monitor environmental pollution [START_REF] Celli | Honey bees as bioindicators of environmental pollution[END_REF][START_REF] Porrini | The death of honey bees and environmental pollution by pesticides: The honey bees as biological indicators[END_REF]. Bumblebees also forage in a great diversity of places and strongly interact with the environment, mainly the flora, surrounding their nests in a range of maximum foraging distances of 363 to 1650 m depending on the species [START_REF] Walther-Hellwig | Foraging distances of Bombus muscorum, Bombus lapidarius, and Bombus terrestris (Hymenoptera, Apidae)[END_REF][START_REF] Wood | Targeted agri-environment schemes significantly improve the population size of common farmland bumblebee species[END_REF], and are thus also suitable organisms for monitoring landscape-based ecological pollution.

While most pesticide research has been focused on managed bees, there has been less work on wild bee populations. For instance, the only European bumblebee that has been studied in relation to pesticide exposure and toxicology is Bombus terrestris, simply because this species is easy to rear in captivity and commercially reared colonies are readily available [START_REF] Baron | Impact of chronic exposure to a pyrethroid pesticide on bumblebees and interactions with a trypanosome parasite[END_REF][START_REF] Gill | Combined pesticide exposure severely affects individual-and colony-level traits in bees[END_REF][START_REF] Rundlöf | Seed coating with a neonicotinoid insecticide negatively affects wild bees[END_REF][START_REF] Whitehorn | Neonicotinoid Pesticide Reduces Bumble Bee Colony Growth and Queen Production[END_REF].

There is increasing evidence that managed bees living in agricultural landscapes are routinely exposed to mixtures of agrochemicals [START_REF] David | Widespread contamination of wildflower and bee-collected pollen with complex mixtures of neonicotinoids and fungicides commonly applied to crops[END_REF][START_REF] Lambert | Widespread occurrence of chemical residues in beehive matrices from apiaries located in different landscapes of Western France[END_REF][START_REF] Long | Non-cultivated plants present a season-long route of pesticide exposure for honey bees[END_REF][START_REF] Mullin | High levels of miticides and agrochemicals in North American apiaries: implications for honey bee health[END_REF][START_REF] Pettis | Crop Pollination Exposes Honey Bees to Pesticides Which Alters Their Susceptibility to the Gut Pathogen Nosema ceranae[END_REF], but little is known about the exposure of wild bees in these environments [START_REF] Hladik | Exposure of native bees foraging in an agricultural landscape to current-use pesticides[END_REF]. Nevertheless, bees constitute a highly diverse group where different taxonomic groups differ widely in their vulnerability to pesticide exposure [START_REF] Biddinger | Comparative toxicities and synergism of apple orchard pesticides to Apis mellifera (L.) and Osmia cornifrons (Radoszkowski)[END_REF][START_REF] Devillers | Comparative toxicity and hazards of pesticides to Apis and non-Apis bees. A chemometrical study[END_REF][START_REF] Piiroinen | Chronic neonicotinoid pesticide exposure and parasite stress differentially affects learning in honeybees and bumblebees[END_REF][START_REF] Thompson | Extrapolation from honeybees to bumblebees in pesticide risk assessment[END_REF]. Furthermore, bee species exhibit pronounced differences in floral preferences and foraging habits, collecting pollen and nectar more frequently from particular plant species according to their morphological traits (e.g. tongue length, body shape and size) and nutritional needs [START_REF] Goulson | Diet breadth, coexistence and rarity in bumblebees[END_REF][START_REF] Vanderplanck | How does pollen chemistry impact development and feeding behaviour of polylectic bees?[END_REF][START_REF] Vaudo | Macronutrient ratios in pollen shape bumble bee (Bombus impatiens) foraging strategies and floral preferences[END_REF].

Such foraging choices may profoundly influence the probability of bees to be more or less A c c e p t e d m a n u s c r i p t exposed to some active compounds [START_REF] Woodcock | Impact of neonicotinoid use on long-term population changes in wild bees in England[END_REF]. Treated crop plants growing in agricultural landscapes have often been regarded as the only source of exposure to agrochemicals for pollinators, but recent research revealed their presence in wild plants growing near crops [START_REF] Botías | Neonicotinoid Residues in Wildflowers, a Potential Route of Chronic Exposure for Bees[END_REF][START_REF] David | Widespread contamination of wildflower and bee-collected pollen with complex mixtures of neonicotinoids and fungicides commonly applied to crops[END_REF][START_REF] Long | Non-cultivated plants present a season-long route of pesticide exposure for honey bees[END_REF][START_REF] Mogren | Neonicotinoid-contaminated pollinator strips adjacent to cropland reduce honey bee nutritional status[END_REF]. We would expect bees that visit flowering arable crops to have higher exposure than those that do not, but also those that visit wild plant species may have varying exposure depending on the ecology, physiology and morphology of their preferred flowers [START_REF] Botías | Contamination of wild plants near neonicotinoid seed-treated crops , and implications for non-target insects[END_REF]. Therefore, it is essential to understand the possible differences in levels of exposure among bee species, since this could reveal which are the most likely exposed and the most frequent mixtures of agrochemicals that they are exposed to.

The widespread occurrence of mixtures of agrochemicals in bee tissues [START_REF] Hladik | Exposure of native bees foraging in an agricultural landscape to current-use pesticides[END_REF] increases concerns regarding the possible detrimental effects of simultaneous exposure to a cocktail of compounds. In general, only the effects of single active substances are studied in toxicity studies both for research and pesticide registration protocols, and exposure to mixtures are only evaluated in risks assessments when they are part of the same formulation.

However, the application of two or more plant protection products during the same cropping season is a common practice in conventional farming [START_REF] Botías | Neonicotinoid Residues in Wildflowers, a Potential Route of Chronic Exposure for Bees[END_REF][START_REF] Garthwaite | Arable crops in the United Kingdom[END_REF], and hence complex mixtures of agrochemicals which are not co-formulants of a single product can be simultaneously detected in bee forage and bee tissues [START_REF] David | Widespread contamination of wildflower and bee-collected pollen with complex mixtures of neonicotinoids and fungicides commonly applied to crops[END_REF][START_REF] Hladik | Exposure of native bees foraging in an agricultural landscape to current-use pesticides[END_REF][START_REF] Long | Non-cultivated plants present a season-long route of pesticide exposure for honey bees[END_REF]. This issue is worrisome given that exposure to mixtures might pose higher risks for animal health than the single impact of a specific class of compounds [START_REF] Cedergreen | Quantifying Synergy: A Systematic Review of Mixture Toxicity Studies within Environmental Toxicology[END_REF][START_REF] Rizzati | Effects of pesticide mixtures in human and animal models: An update of the recent literature[END_REF]. For example, some combinations of insecticides (e.g. pyrethroids with neonicotinoids) and of insecticides with fungicides can lead to additive and synergistic toxicity for bees at the individual and the colony level [START_REF] Gill | Combined pesticide exposure severely affects individual-and colony-level traits in bees[END_REF][START_REF] Iwasa | Mechanism for the differential toxicity of neonicotinoid insecticides in the honey bee, Apis mellifera[END_REF][START_REF] Schmuck | Field relevance of a synergistic effect observed in the laboratory between an EBI fungicide and a chloronicotinyl insecticide in the honeybee (Apis mellifera L, Hymenoptera)[END_REF][START_REF] Sgolastra | Synergistic interactions between pesticides in three bee species[END_REF]. The scarcity of information on the field-relevant mixtures of agrochemicals and levels of exposure for bees could lead us to overlook the possible additive or synergistic effects of pesticide mixtures when risk assessment studies are performed, some of which have been designed to evaluate the hazards of such combinations [START_REF] Sánchez-Bayo | Pesticide residues and bees -a risk assessment[END_REF].

Another major gap in knowledge regarding exposure of bees to pesticides is the potential uptake and contact with these compounds in urban areas, where ornamental nursery plants can also be treated with pesticides [START_REF] Brown | Gardeners Beware : Bee-toxic Pesticides Found in "Bee-Friendly" Plants Sold at Garden Centers Nationwide[END_REF][START_REF] Fevery | Pressure of non-professional use of pesticides on operators, aquatic organisms and bees in Belgium[END_REF] and no information is available on their use in domestic gardens. The possible exposure of bees to harmful A c c e p t e d m a n u s c r i p t pesticides through forage collected in gardens is of high ecological concern, since these habitats are of great value for bees, providing nectar, pollen and nest sites, and sustaining a remarkably high pollinator species richness and abundance [START_REF] Baldock | Where is the UK's pollinator biodiversity? The importance of urban areas for flowervisiting insects[END_REF][START_REF] Kaluza | Urban gardens promote bee foraging over natural habitats and plantations[END_REF][START_REF] Samnegård | Gardens benefit bees and enhance pollination in intensively managed farmland[END_REF]), including bumblebees (Fetridge et al., 2008;[START_REF] Goulson | Effects of land use at a landscape scale on bumblebee nest density and survival[END_REF]. If foraging resources and nesting sites in urban habitats are contaminated with pesticide residues, it is likely that exposure to certain active compounds could be more widely spread in the landscape and more prolonged in time than previously assumed.

The aim of this study was to evaluate and compare exposure in different wild bumblebee species. To do this, we analysed the levels of five neonicotinoid insecticides, thirteen currentlyused fungicides and a pesticide synergist in tissues of five bumblebee species (B. hortorum, B. pascuorum, B. terrestris, B. lapidarius and B. pratorum). These wild bumblebee samples were collected in agricultural and urban habitats to compare levels of exposure in both environments and to study distribution of agrochemicals in the landscape. The bees were gathered in three different periods (late spring, early summer and midsummer) in order to monitor the length of exposure to agrochemicals through the season.

Our results show evidence that wild bumblebees are frequently exposed to mixtures of agrochemicals when they forage in arable and urban habitats, with peak concentrations decreasing through the season.

Materials and Methods

Sampling sites and field collection

Wild bumblebees were collected in five farms and five urban landscapes in East Sussex (South-East England, UK), all sites being at least 2 kilometers apart from each other (Figure 1) (Table S1). The sites selected to collect bees in agricultural land consisted of arable fields within mixed farms, where the predominant crops were oilseed rape, winter wheat and spring barley, and part of the land was pasture. The urban sampling sites consisted of ornamental public gardens and parks surrounded by houses that had private gardens in most cases. Foraging bumblebees were collected using insect nets and kept in individual labeled tubes and put on ice during transport back to the lab, and then kept at -80°C until pesticide analysis was performed. Specimens of five bumblebee species with different ranges of tongue length were sampled [START_REF] Brodie | Bumblebee foraging preferences: differences between species and individuals[END_REF][START_REF] Prys-Jones | Bumblebees[END_REF]): short-tongued bumblebees were B. pratorum (6.4 -7.1 mm), B. lapidarius (6 -8.1 mm), B. terrestris (5.8 -8.2 mm); mediumtongued was B. pascuorum (7.6 -8.6 mm); and long-tongued was B. hortorum (12 -13.5 mm) A c c e p t e d m a n u s c r i p t (Table 1). The flowers where the bees were foraging at the time of capture were recorded (Tables S2a-S1c), since bumblebees exhibit a high degree of floral constancy [START_REF] Wilson | Floral Constancy in Bumble Bees : Handling Efficiency or Perceptual Conditioning ?[END_REF], and this may help predict exposure. Bumblebee individuals were gathered during three sampling periods, spring (27/04/14 -14/05/14), early summer (5/06/14 -23/06/14) and midsummer (15/07/14 -2/08/14), and 150 bee individuals were collected in total. Oilseed rape crops were in bloom during the first sampling period (late spring), and 18 out of the 25 individuals gathered in arable sites during that period were foraging in oilseed rape crops when collected (Table S2a). The pesticide usage information of the crops where bees were foraging was not provided by the farmers. The EU moratorium on the use of neonicotinoid insecticides started on the 1 st December 2013, but the oilseed rape crops that were in bloom in the 2014 spring were sown at the end of Augustbeginning of September 2013, so these crops were still allowed to be seed-treated with neonicotinoids.

Pesticide analysis

Chemicals and reagents

Eight classes of contaminants were chosen to be tested in the bumblebee samples, including the five neonicotinoid insecticides that are registered for use in the UK, based on the most used (by weight or area treated) in UK crops including oilseed rape, wheat, spring barley, field bean, strawberry and raspberry crops (https://secure.fera.defra.gov.uk/pusstats/surveys/2014surveys.cfm) (Table S3). Except for certified standards of carbendazim-d3 and tebuconazole-d6, which were purchased from LGC standards UK and carbamazepine-d10 and prochloraz-d7 from QMX Laboratories Limited UK, all the other certified standards as well as formic acid, magnesium sulphate, ammonium formate, sodium acetate and Supel TM QuE PSA/C18/GCB (ratio 1/1/1) were obtained from Sigma Aldrich UK. The compound purity of the pesticide standards was > 99%, apart from spiroxamine (98.5%), triticonazole (98.8%), piperonyl butoxide (97.9%). Isotopic purity of deuterated standards was > 97%. HPLC grade acetonitrile, toluene and water were obtained from Rathburns UK.

Preparation of samples and residue analysis

The extraction of pesticides from the bumblebee samples was performed as reported in [START_REF] David | Sensitive determination of mixtures of neonicotinoid and fungicide residues in pollen and single bumblebees using a scaled down QuEChERS method for exposure assessment[END_REF]. (QuEChERS method). After the dSPE step, the sorbent was extracted with acetonitrile/toluene (3/1, 150 µl) and the supernatant was combined with that of the previous acetonitrile extract and spin filtered (0.22 µm). After evaporation, reconstitution was made with 120 µl acetonitrile:water (30:70) and the extract was centrifuged (20 min). The supernatant obtained was stored in the dark at -20°C before analysis.

UHPLC-MS/MS analyses

Extracts of samples were analysed by ultra high-performance liquid chromatography tandem mass spectrometry (UHPLC-MS/MS) as described in [START_REF] David | Sensitive determination of mixtures of neonicotinoid and fungicide residues in pollen and single bumblebees using a scaled down QuEChERS method for exposure assessment[END_REF], using a Waters Acquity UHPLC system coupled to a Quattro Premier triple quadrupole mass spectrometer from Micromass (Waters, Manchester, UK). Acquisition of data was performed with the software MassLynx 4.1 and the quantification was established by the calculation of the response factor of each pesticide to its internal standard. Least-square linear regression analysis of the native analyte to deuterated internal standard (concentration ratio) versus the peak area was used to determine analyte concentrations. Further methodological details are described in [START_REF] David | Sensitive determination of mixtures of neonicotinoid and fungicide residues in pollen and single bumblebees using a scaled down QuEChERS method for exposure assessment[END_REF].

Table S4 reports the method detection limits (MDL) and the method quantification limits (MQL) for all the compounds that were analysed in bumblebees.

Statistical analysis

Shapiro-Wilk test was used to assess whether data (i.e. concentration of pesticides detected in the bumblebees per species, habitat, sampling period and bumblebee body mass) met the assumptions of normality. Since pesticide concentrations did not meet the assumptions of normality, non-parametric Kruskal-Wallis tests were used to compare this variable among the 5 different bee species and 3 sampling periods. Pairwise comparisons between bee species, sampling periods and habitats were performed using Mann-Whitney U-tests. Bumblebee body mass and frequencies of pesticide detection in the 5 species were normally distributed, so oneway ANOVA analyses were carried out to compare weights of the 5 species, using Bonferroni For the statistical analysis, the MDL value was assigned to concentrations above the limits of detection (≥MDL) but below the limits of quantification (>MQL), whereas those below the MDL were considered to be zero. The software used to perform the statistical analysis was SPSS 21.

Results

Ranges, frequencies and average levels of neonicotinoid and fungicide residues detected in wild bumblebee samples

In total, 60.7% of the 150 individuals tested had detectable levels of at least one of the compounds analysed, with boscalid being the most frequently detected (in 35.3 % of all the bees analysed), followed by tebuconazole (27.3%), spiroxamine (18.7%), carbendazim (10.7%), epoxiconazole (8%), imidacloprid (7.3%), metconazole (6.7%) and thiamethoxam (6%) (Table 2) (Tables S2a-2c). In general, from all the bees where pesticides were detected, the majority (71.4%) had more than one compound, and 43.4 % of the bees contained two or more, with a maximum of 7 different pesticides being detected in one specimen (B. lapidarius collected from urban site 2 in June) (Table S2b).

Levels of pesticide exposure for the five bumblebee species studied

Although the combinations of pesticides detected varied notably among individuals of the same and the other species, some combinations occurred very frequently in farmland bumblebees. Boscalid with DMI-fungicides was the most common mixture detected in bumblebees (64%) foraging in farmland in spring (April-May), except for B. pratorum individuals where such combination was not detected (Table S2a). In early summer (June) spiroxamine with DMI-fungicides was detected in the great majority of farmland bumblebees (88%), regardless of the species (Table S2b) (Figures 2 and3). In general, the detection frequency of the 19 agrochemicals analysed did not significantly differ among the 5 species (ANOVA, F (4, 25) = 0.78, P = 0.55). However, when the concentrations of pesticides detected in the 5 bumblebee species were compared, we found that B. pratorum, the species with the lowest body mass and shortest tongue length range (

A c c e p t e d m a n u s c r i p t

In order to evaluate the relationship between bee body size and levels of exposure, we compared the body mass of the five bumblebee species, and we found that B. pratorum was significantly lighter than the two species with the highest pesticide concentrations, B. terrestris

(1-way ANOVA with Bonferroni´s multiple comparison test: P = 0.03) and B. lapidarius (P = 0.017).

Levels of pesticide exposure in arable and urban habitats

In general, bees foraging in agricultural landscapes had significantly higher levels of agrochemicals (6.8 ± 9.5 ng/g) than those foraging in urban sites (2.5 ± 7.8 ng/g)(M-W test:

U(148) = 1635.5; Z = -4.6; P < 0.001)(Figure 2). However, the highest levels and frequencies of detection for neonicotinoids (10 ng/g of imidacloprid on a B. terrestris specimen) and the most frequently detected fungicide boscalid (54.5 ng/g in a B. lapidarius specimen) were recorded in urban bumblebees collected during the early summer (June) (Table S2b). Overall, neonicotinoids were found in more bees in urban sites than in farmland (9.3% versus 2.7%), with all five neonicotinoids registered for use in the UK found in at least one urban bee.

Changing levels of pesticide exposure through the season

The levels of exposure to agrochemicals for wild bumblebees were examined for the period of highest foraging activity in the studied area (East Sussex, England), and we found that the frequencies of detection decreased both in arable and urban habitats for the 5 species evaluated in midsummer (July) (Figure 3), when only 28% of the bees collected had at least one agrochemical, compared to 76% in late spring (April-May) and 78% in early summer (June).

Consequently, the average concentrations detected were lower in midsummer (July: 0.6 ± 2.3 ng/g) than in spring (April-May: 5.9 ± 7.6 ng/g) (M-W test: U(98) = 474; Z = -5.67; P < 0.001) and early summer (June: 7.5 ± 12.4 ng/g) (M-W test: U(98) = 462.5; Z = -5.74; P < 0.001) (Table 2).

Discussion

Our field study revealed that free-flying wild bumblebees are exposed to multiple pesticide residues, with different levels and frequencies of detection according to the species, sampling period and landscape context. Several studies have reported the presence of mixtures of agrochemicals in honeybee matrices [START_REF] Lambert | Widespread occurrence of chemical residues in beehive matrices from apiaries located in different landscapes of Western France[END_REF][START_REF] Mullin | High levels of miticides and agrochemicals in North American apiaries: implications for honey bee health[END_REF][START_REF] Pettis | Crop Pollination Exposes Honey Bees to Pesticides Which Alters Their Susceptibility to the Gut Pathogen Nosema ceranae[END_REF], where more than 170 compounds have been detected so far (Sánchez-Bayo and Goka, 2014), but little is known about the exposure of wild bees. Recent research detected up to 19

A c c e p t e d m a n u s c r i p t different chemicals in native bees collected from wheat fields and grassland in Colorado (USA) [START_REF] Hladik | Exposure of native bees foraging in an agricultural landscape to current-use pesticides[END_REF], reporting levels as high as 310 ng/g for thiamethoxam, 87 ng/g for clothianidin and 82 ng/g for imidacloprid. The maximum concentrations detected in our bumblebee samples were much lower (i.e. 2.35 ng/g for thiamethoxam, 1.4 ng/g for clothianidin and 10 ng/g for imidacloprid), which could be explained by the differences between both experimental designs. While we collected free-flying bumblebees with nets and analysed them individually, [START_REF] Hladik | Exposure of native bees foraging in an agricultural landscape to current-use pesticides[END_REF] performed the analysis on composite samples containing approximately 10 individuals of different wild bee species that were collected using bee monitoring traps. The collection and analysis of composite samples containing individuals from different bee genera might conceivably increase the chances of including particular specimens that could have been exposed to very high concentrations of pesticides due to their foraging and feeding behaviour, metabolic rates and morphological traits. Although the bumblebee species analysed in our study can present differences in their foraging distances, they seem to use and benefit similarly from the resources available in farmland [START_REF] Wood | Targeted agri-environment schemes significantly improve the population size of common farmland bumblebee species[END_REF], while bees from other genera, specially solitary bees, may present more variation in their foraging choices [START_REF] Wood | Providing foraging resources for solitary bees on farmland: current schemes for pollinators benefit a limited suite of species[END_REF], and thus, a wider range of levels of exposure to agrochemicals. On the other hand, the dissimilarities in pesticide use patterns between the UK and the USA could partly explain the differences found, as for instance, the maximum application rate for thiamethoxam in oil seed crops in the UK is 33.6 g a.i./ha (European Food Safety Authority, 2013), and 157 g a.i./ha in the USA (USEPA, 2011). Studying the link between pesticide application rates and the levels of exposure for bees, and how different bee species can be more or less susceptible to exposure is essential for a full understanding of the risk posed by pesticides.

The comparison of pesticide concentrations among the five bumblebee species that we studied showed that B. pratorum, the species with the smallest body mass and tongue length range, had lower residue levels than the other four species. Different explanations are plausible; for example smaller bees may consume lower amounts of food, and hence they would be less exposed to these active compounds present in pollen and nectar. Smaller body size may lead to greater mass-specific metabolic rates [START_REF] Heinrich | The hot-blooded insects. Strategies and mechanisms of thermoregulation[END_REF], and so pesticides might be metabolised faster in smaller bees such as B. pratorum, whose body mass was significantly lower than that of the two species with the highest pesticide concentrations, B. terrestris and

B. lapidarius.

A c c e p t e d m a n u s c r i p t

Neonicotinoid insecticides can be metabolised relatively fast, with metabolites being the main residues detected in bees a few minutes after ingestion of the parent compound [START_REF] Brunet | In vivo metabolic fate of [14C]-acetamiprid in six biological compartments of the honeybee, Apis mellifera[END_REF][START_REF] Suchail | Metabolism of imidacloprid in Apis mellifera[END_REF]. Also the metabolites of some fungicides have been detected in bees and bee-collected pollen [START_REF] Jabot | Detection and quantification of boscalid and its metabolites in honeybees[END_REF][START_REF] Mullin | High levels of miticides and agrochemicals in North American apiaries: implications for honey bee health[END_REF][START_REF] Stoner | Using a Hazard Quotient to Evaluate Pesticide Residues Detected in Pollen Trapped from Honey Bees (Apis mellifera) in Connecticut[END_REF], so the analysis of neonicotinoid and fungicide metabolites could have revealed the presence of other potentially toxic compounds that bees might have been exposed to. Some neonicotinoid metabolites have been proven to be highly toxic for bees [START_REF] Simon-Delso | Systemic insecticides (neonicotinoids and fipronil): trends, uses, mode of action and metabolites[END_REF][START_REF] Suchail | Metabolism of imidacloprid in Apis mellifera[END_REF][START_REF] Suchail | Discrepancy between acute and chronic toxicity induced by imidacloprid and its metabolites in Apis mellifera[END_REF], while the possible effects of both parent fungicides and their metabolites has been scarcely studied in bees, with some reports showing detrimental effects [START_REF] Bernauer | Colonies of Bumble Bees (Bombus impatiens) Produce Fewer Workers, Less Bee Biomass, and Have Smaller Mother Queens Following Fungicide Exposure[END_REF][START_REF] Pettis | Crop Pollination Exposes Honey Bees to Pesticides Which Alters Their Susceptibility to the Gut Pathogen Nosema ceranae[END_REF][START_REF] Syromyatnikov | Evaluation of the toxicity of fungicides to flight muscle mitochondria of bumblebee (Bombus terrestris L.)[END_REF][START_REF] Vanengelsdorp | Entombed Pollen": A new condition in honey bee colonies associated with increased risk of colony mortality[END_REF]. It is possible that the action of fungicides on bees may not be directly toxic, as is the case with insecticides, but may alter the beneficial microbiome present in the pollen and nectar [START_REF] Bartlewicz | Effects of agricultural fungicides on microorganisms associated with floral nectar: susceptibility assays and field experiments[END_REF][START_REF] Vanengelsdorp | Entombed Pollen": A new condition in honey bee colonies associated with increased risk of colony mortality[END_REF][START_REF] Yoder | Fungicide contamination reduces beneficial fungi in bee bread based on an area-wide field study in honey bee, Apis mellifera, colonies[END_REF] and as a consequence, in the bee gut, which could have important implications for bee nutrition and health [START_REF] Engel | The Bee Microbiome: Impact on Bee Health and Model for Evolution and Ecology of Host-Microbe Interactions[END_REF][START_REF] Mattila | Characterization of the active microbiotas associated with honey bees reveals healthier and broader communities when colonies are genetically diverse[END_REF][START_REF] Pettis | Crop Pollination Exposes Honey Bees to Pesticides Which Alters Their Susceptibility to the Gut Pathogen Nosema ceranae[END_REF].

The tongue length of the different bumblebee species could be hypothesized as a possible predictor of residue exposure, since this trait determines whether or not, and how fast, a bee can manipulate a particular flower to extract nectar, and thus it is crucial for the division of resources between different species [START_REF] Brian | Differences in the Flowers Visited by Four Species of Bumble-Bees and their Causes[END_REF][START_REF] Cariveau | The Allometry of Bee Proboscis Length and Its Uses in Ecology[END_REF][START_REF] Goulson | Diet breadth, coexistence and rarity in bumblebees[END_REF][START_REF] Harder | Morphology as a Predictor of Flower Choice by Bumble Bees[END_REF]. Long-tongued bees generally forage from long-corolla flowers, and shorttongued bees from short corolla flowers [START_REF] Hobbs | Further studies on the food-gathering behaviour of bumble bees (Hymenoptera: Apidae)[END_REF]. In the case of B. pratorum, as a shorttongued bee, the most commonly visited flowers should be those with short corolla (e.g. many Rosaceae and Asteraceae flowers), which have both nectaries and stamens more exposed to environmental conditions and wind-blown aerosols. Oilseed rape flowers are shallow and more frequently visited by short-tongued bees, even though the long-tongued bumblebee B.

hortorum often collects pollen from this plant [START_REF] Stanley | Pollinators and pollination of oilseed rape crops (Brassica napus L.) in Ireland: Ecological and economic incentives for pollinator conservation[END_REF], and all our B. hortorum specimens sampled in farmland in late spring were foraging in these crop flowers when collected (Table S2a). Some of the pesticides analysed in this study have been reported to degrade after exposure to sun light and/or high temperatures (i.e. some neonicotinoids, carbendazim, carboxin, epoxiconazole and prochloraz) [START_REF] Bonmatin | Environmental fate and exposure; neonicotinoids and fipronil[END_REF][START_REF] Burrows | Reaction pathways and mechanisms of photodegradation of pesticides[END_REF][START_REF] Mazellier | Photochemical behavior of the fungicide carbendazim in dilute aqueous solution[END_REF], so it is possible that the flowers with pollen and nectar more exposed to the environmental conditions might have lower concentrations of the parent active compounds. Therefore, the bees feeding on these shallow flowers would be less exposed to them. However, the tongue length range of B. pratorum is not very different to that of the A c c e p t e d m a n u s c r i p t short-tongued bumblebees B. terrestris and B. lapidarius (Table 1). Moreover, the range of flowers visited by the three species did not differ remarkably, since more than half (53%) of the plant species visited by B. terrestris and B. lapidarius were also visited by B. pratorum in May and June (i.e. when concentrations detected were higher) (Tables S2a-2b). Nevertheless, as mentioned above, B. terrestris and B. lapidarius showed significantly higher levels of pesticide concentrations than B. pratorum, so the tongue length doesn´t seem to be a suitable predictor of residue exposure for the group of bumblebees species studied here. Otherwise, a bigger sample size might be needed to test this hypothesis.

Regarding the toxicity of the pesticides detected, it is worth noting that the bumblebee specimens collected in the present study were individually processed as whole samples to include residues on external as well as internal parts of the bees, so it is not possible to differentiate if the pesticides detected were on the cuticle (contact toxicity) or inside the organism (oral). Thus, both routes of exposure should be considered when the levels detected in the samples are compared to lethal doses reported for the compounds analysed. Moreover, as the bee gut was not removed before processing the samples, there is a chance that some of the residues detected were present in the nectar and pollen contained in the digestive tract, although we consider that the bulk of the bee weights were formed by bee tissues. None of the residues detected in bumblebees were found to overlap with contact or oral acute LD 50 values tested on bumblebees or honeybees (Table 2), which is to be expected since the bees screened for pesticides in this study were performing foraging tasks and appeared to be healthy at the time of collection; it would be very unlikely to catch bees alive had they been exposed to lethal doses. Additionally, we cannot determine what doses the bees had been exposed to since pesticides are metabolized at varying rates (and we do not know the time of exposure), so that the residues we detected represent an unknown proportion of the dose received and actual exposures may have probably been higher. It should also be mentioned that bees are subjected to chronic exposure when foraging on contaminated flowers, and acute LD 50 s are frequently higher than chronic LD 50 s, particularly for neonicotinoids [START_REF] Alkassab | Impacts of chronic sublethal exposure to clothianidin on winter honeybees[END_REF][START_REF] Rondeau | Delayed and time-cumulative toxicity of imidacloprid in bees, ants and termites[END_REF]. Thus, the potential risk of chronic exposure to the levels of pesticides detected in the bumblebees cannot be ruled out. Studies where pesticide mixtures are analysed in bees showing health problems or mortality in the field would be of high interest [START_REF] Kasiotis | Pesticide residues in honeybees, honey and bee pollen by LC-MS/MS screening: Reported death incidents in honeybees[END_REF], since this could provide key information about the hazard posed by specific mixtures and the threshold levels of risk. Nevertheless, such studies are highly challenging to perform in the field due to sampling difficulties, and because the detection of accurate levels at the time of death are problematic since bee samples need to be A c c e p t e d m a n u s c r i p t fresh to avoid degradation of pesticides following exposure to environmental factors and microbial activity [START_REF] Katagi | Photodegradation of pesticides on plant and soil surfaces[END_REF][START_REF] Liu | Soil microbial degradation of neonicotinoid insecticides imidacloprid, acetamiprid, thiacloprid and imidaclothiz and its effect on the persistence of bioefficacy against horsebean aphid Aphis craccivora Koch after soil application[END_REF].

Our results revealed that 43.3% of bees contained two or more pesticides, suggesting that simultaneous exposure is likely to be encountered regularly in a field realistic scenario.

Although contamination with mixtures of pesticides has been detected in flower pollen, bee collected pollen and bees [START_REF] David | Widespread contamination of wildflower and bee-collected pollen with complex mixtures of neonicotinoids and fungicides commonly applied to crops[END_REF][START_REF] Hladik | Exposure of native bees foraging in an agricultural landscape to current-use pesticides[END_REF][START_REF] Long | Non-cultivated plants present a season-long route of pesticide exposure for honey bees[END_REF][START_REF] Mullin | High levels of miticides and agrochemicals in North American apiaries: implications for honey bee health[END_REF], evaluating the impact of the exposure to such combinations of agrochemicals poses a major challenge and warrants more research. A few laboratory and field studies have explored the impact of simultaneous exposure to different chemicals on bees, some of them reporting detrimental effects of certain combinations [START_REF] Gill | Combined pesticide exposure severely affects individual-and colony-level traits in bees[END_REF][START_REF] Iwasa | Mechanism for the differential toxicity of neonicotinoid insecticides in the honey bee, Apis mellifera[END_REF][START_REF] Johnson | Acaricide, fungicide and drug interactions in honey bees (Apis mellifera)[END_REF][START_REF] Schmuck | Field relevance of a synergistic effect observed in the laboratory between an EBI fungicide and a chloronicotinyl insecticide in the honeybee (Apis mellifera L, Hymenoptera)[END_REF][START_REF] Sgolastra | Synergistic interactions between pesticides in three bee species[END_REF][START_REF] Thompson | Assessment of the synergy and repellency of pyrethroid / fungicide mixtures[END_REF]. Given the mixtures detected in bees in the present study and in previous research [START_REF] Hladik | Exposure of native bees foraging in an agricultural landscape to current-use pesticides[END_REF][START_REF] Lambert | Widespread occurrence of chemical residues in beehive matrices from apiaries located in different landscapes of Western France[END_REF], the number of possible combinations of pesticides is very high and variable. Therefore, focussing on the most frequent ones when performing assays of bee exposure might be the most sensible approach. Moreover, these scenarios are especially important to consider in cases when two or more pesticides that exhibit synergy are detected simultaneously in bees. For instance, the toxicity of certain insecticides (e.g. neonicotinoids and pyrethroids) can be enhanced in the presence of demethylation-inhibiting (DMI) fungicides (e.g. epoxiconazole, tebuconazole). In our study, 55.6 % of the bumblebees where neonicotinoid insecticides were detected also contained DMI-fungicides, so exposure to these combinations seems to be likely in the field although it is not known if these concentrations were high enough to induce biological effects. These DMI-fungicides can act as synergists by inhibiting the detoxification system in bees [START_REF] Iwasa | Mechanism for the differential toxicity of neonicotinoid insecticides in the honey bee, Apis mellifera[END_REF][START_REF] Johnson | Acaricide, fungicide and drug interactions in honey bees (Apis mellifera)[END_REF][START_REF] Pilling | Mechanism of Synergism between the Pyrethroid Insecticide alpha-Cyhalothrin and the Imidazole Fungicide Prochloraz in the Honeybee (Apis mellifera L.)[END_REF], and thus the insecticide residues are metabolised or eliminated more slowly. It is also important to remark that this is a limited list of pesticides; due to analytical constrains, insecticides such as pyrethroids, usually detected using gas chromatographic methods and which are known to interact with neonicotinoids and DMI-fungicides, could also be present in these bees.

As for the differences in levels of exposure for bees living in farmland or urban habitats, we found that concentrations of pesticides in wild bumblebees foraging in agricultural land were higher than in urban land, as reported for commercial bumblebee colonies in a previous study [START_REF] David | Widespread contamination of wildflower and bee-collected pollen with complex mixtures of neonicotinoids and fungicides commonly applied to crops[END_REF]. However, the maximum values for neonicotinoids were recorded in bumblebees collected in urban gardens (i.e. 10 ng/g of imidacloprid, 2.35 ng/g of A c c e p t e d m a n u s c r i p t thiamethoxam and 1.4 ng/g of clothianidin), even though these maximum values were high in comparison to the levels detected for these compounds in the rest of the bee specimens collected, so it is possible that these particular bees had been visiting freshly sprayed plants just before they were sampled. The use of imidacloprid, clothianidin and thiamethoxam has been banned since December 2013 on ornamental plants flowering in the year of treatment (as well as on flowering crops) (European Commission, 2013), and so the high levels of imidacloprid in particular are hard to explain. Nonetheless, the result is corroborated by [START_REF] David | Widespread contamination of wildflower and bee-collected pollen with complex mixtures of neonicotinoids and fungicides commonly applied to crops[END_REF] who detected high levels (20 ng/g) of imidacloprid in pollen collected by commercial bumblebee colonies placed in urban areas. The imidacloprid may be persisting in urban environments from applications before the ban, or it might still be available in some stores and presumably may be used in gardens for months or even years after the ban.

Alternatively, this exposure of bees could originate from other uses of imidacloprid; it is widely used in baits to kill ants and for flea control in dogs and cats. Orally applied imidacloprid in dogs and other mammals is completely eliminated in the urine (70-80%) and faeces (20-30%) in 48 h as the main metabolites 6-chloronicotinic acid and its glycine conjugate together with significant amounts of the parent compound, whereas topical application spreads over the skin for 24 hours and the compound is stored in the oil glands of the skin and slowly shed with hair and sebum (European Food Safety Authority, 2006;[START_REF] Hovda | Toxicology of newer pesticides for use in dogs and cats[END_REF]. The environmental impacts of the use of pesticides in ornamental gardens and on pets has been scarcely studied [START_REF] Brown | Gardeners Beware : Bee-toxic Pesticides Found in "Bee-Friendly" Plants Sold at Garden Centers Nationwide[END_REF][START_REF] Fevery | Pressure of non-professional use of pesticides on operators, aquatic organisms and bees in Belgium[END_REF]. There is no policing of homeowner use of garden pesticides to ensure that they follow label instructions, and amounts used in gardens are not known. Similarly, we can find no information on the number of dogs and cats treated with imidacloprid as a prophylactic flea treatment, and the environmental fate of such chemicals has not been studied. Urban gardens are an important food source and refuge for bees and other wildlife in cities and towns because they represent the only green space in these large environments, and they can host a great diversity of pollinators and high density of bumblebee nests [START_REF] Baldock | Where is the UK's pollinator biodiversity? The importance of urban areas for flowervisiting insects[END_REF]Fetridge et al., 2008;[START_REF] Goulson | Effects of land use at a landscape scale on bumblebee nest density and survival[END_REF], resulting in enhanced pollination services for both urban and nearby agricultural landscapes [START_REF] Kaluza | Urban gardens promote bee foraging over natural habitats and plantations[END_REF][START_REF] Samnegård | Gardens benefit bees and enhance pollination in intensively managed farmland[END_REF][START_REF] Theodorou | Pollination services enhanced with urbanisation despite increasing pollinator parasitism[END_REF]. If bees are sometimes exposed to high doses of harmful pesticides through forage collected in ornamental gardens, this might be a matter of high ecological concern, meaning that more attention should be paid to this route of contamination for bees.

The frequencies of detection and concentrations of agrochemicals in bee tissues decreased significantly in midsummer (July), agreeing with findings reported before [START_REF] Botías | Neonicotinoid Residues in Wildflowers, a Potential Route of Chronic Exposure for Bees[END_REF] A c c e p t e d m a n u s c r i p t [START_REF] David | Widespread contamination of wildflower and bee-collected pollen with complex mixtures of neonicotinoids and fungicides commonly applied to crops[END_REF] where residue levels decreased in the pollen collected by bees as the season progressed. Midsummer is usually the period for crop harvesting in the studied area, and fewer pesticides are normally applied during this period to crops, so this may partly explain this finding. Concurrently, the decrease in residue levels may be due to a reduction in plant tissue concentrations, and thus on pollen and nectar collected by bees through summer because of plant/soil metabolism, photolysis and increasing temperatures [START_REF] Bonmatin | Environmental fate and exposure; neonicotinoids and fipronil[END_REF][START_REF] Lassalle | UV-visible degradation of boscalid -Structural characterization of photoproducts and potential toxicity using in silico tests[END_REF][START_REF] Mazellier | Photochemical behavior of the fungicide carbendazim in dilute aqueous solution[END_REF].

Conclusions and perspectives

The extensive incidence of multiple pesticide residues and the scarcity of scientific literature on the biological consequences of exposure to such combinations on wild bees calls for more attention on regulatory strategies concerning monitoring procedures and registration of agrochemicals as they relate to pollinator protection, since the combined toxicity and synergism of all these chemicals may pose a real threat to the health and survival of managed and wild bees. Thus, investigations on the toxicological effects of field-realistic levels and mixtures are crucial to prevent potential exposure of bees to damaging combinations of agrochemicals. Also, the possible impact of metabolites and of agrochemicals other than insecticides on bees should not be overlooked, as they could have direct or indirect detrimental effects.

Understanding the factors involved in the degree of exposure, such as the type of flowers that tend to incorporate more residues or where certain pesticides are more persistent, as well as the morphological and physiological traits of pollinators that makes them more or less susceptible to the exposure, is also crucial to mitigate the damaging effects for the most vulnerable species.

Finally, the widespread detection of pesticides in bumblebees foraging in urban areas indicates a pressing need for further research on the prevalence and doses present in ornamental gardens, and on the environmental fate of pesticides upon domestic uses in order to better inform homeowners and garden centers of the potential risk the use of these products poses.

Furthermore, surveillance programs on domestic uses would improve the current lack of safety and control in the application of agrochemicals in non-commercial agricultural situations.

Our results show evidence that wild bumblebees are frequently exposed to mixtures of agrochemicals when they forage in arable and urban habitats, with peak concentrations decreasing through the season. The effects of exposure to pesticide mixtures in wild bees

  whole bumblebee specimens (N = 150) were ground in liquid nitrogen and weighed (mean weight ± S.D = 108 ± 46 mg; range = 21 -236 mg). Each sample was spiked with 10 µl acetonitrile containing the mixture of deuterated internal standards (IS) at 40 ng/ml (400 pg of each IS). Subsequently, the extraction was performed by the addition of 400 µl of water, 500 µl of acetonitrile, 250 mg of magnesium sulphate: sodium acetate mix (4:1) and 50 mg of SupelTM QuE PSA/C18/GCB for the dispersive solid phase extraction step (dSPE)

  tests for pairwise comparisons, and to compare the frequencies of detection among them.

  determined, but studying the temporal distribution of such combinations in habitats favored by bees is crucial to identify timing and routes of pesticide exposure, which may help us to properly direct our conservation efforts regarding pesticide regulation and bee health protection.

Figure 1 .

 1 Figure1. Location of sites in East Sussex (UK), where wild bumblebees were collected from farmland sites (circles) and urban sites (squares). The sites selected to collect bees in agricultural land consisted of arable fields within mixed farms, where the predominant crops were oilseed rape, winter wheat and spring barley, and part of the land was pasture. The urban sampling sites consisted of ornamental public gardens and parks surrounded by houses that had private gardens in most cases.

Figure 2 .

 2 Figure 2. Sum of total concentrations of neonicotinoid insecticides and fungicides detected in 5 different species of wild bumblebees collected in arable (A) and urban (U) sites in three sampling periods (May, June and July). Concentrations reported refer to the amount of active substance (ng) detected per bumblebee body weight (g).

Figure 3 .

 3 Figure 3. Frequencies of detection (%) of pesticide residues in wild bumblebees collected in arable and urban habitats in May, June and July.

Table 1

 1

	), had significantly lower residue

P = 0.013), B. terrestris (6.8 ± 10.4 ng/g) (U(58) = 275; P = 0.006), B. lapidarius (7.2 ± 11.8) (U(58) = 260; P = 0.003) and also tended to have lower concentrations than B. pascuorum (2.8 ± 4.9) (U(58) = 330; P = 0.056).

Table 1 .

 1 Mean (± standard deviation) and range of body mass values (mg), and tongue length range (mm) for the five bumblebee species analysed.

	BUMBLEBEE	BODY MASS (mg)	TONGUE LENGTH RANGE (mm)
	SPECIES	MEAN ± S.D.	RANGE	(Brodie, 1996; Pry-Jones and Corbet, 2011)
	B. hortorum	105 ± 45	40 -223	12 -13.5
	B. pascuorum	97 ± 34	29 -171	7.6 -8.6
	B. terrestris	142 ± 46	45 -236	5.8 -8.2
	B. lapidarius	117 ± 49	40 -226	6 -8.1
	B. pratorum	79 ± 32	21 -161	6.4 -7.1

Table 2 .

 2 Frequencies of detection (%), maximum concentrations (Max), average (Avg) and median values (Mdn)(ng/g) of neonicotinoid insecticides and fungicides detected in wild bumblebees. The analytical methods do not allow us to differentiate what fraction of the pesticide was on the surface (contact toxicity) or inside the bumblebee (oral) since all

	specimens were individually processed as whole samples. Therefore, LD 50 values of contact (C)
	and/or oral (O) toxicity for honeybees (hb) or bumblebees (bb), according to availability of
	data (Sánchez-Bayo and Goka, 2014), are reported for each of the pesticides that were
	detected in wild bumblebees.

Acknowledgements

We are grateful to Defra (Research Project PS2372) for funding this work and to the five farmers for allowing us to work on their property. We are also grateful to the Sheepdrove Trust for contributing to the costs of the analytical work.