In present study, the classification of our hydra IMBE1 clone has been reevaluated to dispel doubt concerning its species name. At the time of writing, this clone has been in breeding for at least the past four decades. In the early 1980’s, it was used at the University of Philadelphia (USA) under the name Hydra attenuata in a prescreening test for the detection of substances with teratogenic potential in mammals (e.g. Johnson et al. 1982). The same clone was later used also by Environment Canada, and in France, in the Universities of Lorraine and Aix Marseille in standardized assays to assess the toxicity of polluted freshwaters and wastewaters as well as pure chemical compounds (Blaise and Kusui 1997; Blaise et al. 2018; Colpaert et al. 2020; De Jong et al. 2016; Pachura et al. 2005; Pachura-Bouchet et al. 2006; Quinn et al. 2008). Due to the confusion created by Schulze in 1917 around the name « attenuata » which has been used both to describe specimens belonging to the Hydra vulgaris and Hydra circumcincta species (Campbell 1989), the clones used in the previously cited works were either named as Hydra attenuata or Hydra circumcincta. The phylogenetic tree generated for this study clearly showed that the closest relatives of our IMBE1 clone, which had travelled between North America and Europe, belonged to the Hydra vulgaris species and were of European origin suggesting that it had itself the same geographic origin.
The outstanding regeneration capacity of species of Hydra genus has been successfully used in ecotoxicology by several authors to detect the teratogenic potential of numerous chemicals. To our knowledge, the pioneer was Johnson (1980) who used pellets of hydra dissociated cells as ‘artificial embryo’. Interestingly, using eight pharmaceuticals already evaluated in mammals (dexamethasone, aspirin, retinol acetate, methotrexate, actinomycin D, vinblastine and non-acetylated isoniazid), he demonstrated that the low expensive hydra assays both performed on entire adult polyps and on ‘artificial embryos’ could be useful to identify non-coaffective and coaffective teratogens as adequately than the expensive mammalian assays on rodents. In subsequent studies (Johnson et al. 1982; Johnson and Gabel 1983), Johnson and collaborators further expanded their original findings to a greater family of compounds including formaldehyde, benzene, lithium carbonate, lithium chloride, DMSO, flavor enhancers, food additives (e.g. monosodium glutamate), phenylenediamine (hair dye product), saccharine, colchicine as well as several phthalates (dimethyl phthalate, diethyl phthalate, dibutyl phthalate) which are known endocrine disruptors (EDs) (e.g. Mathieu-Denoncourt, 2015 for review). For all these compounds, Johnson and collaborators found a good correlation between the A/D ratios observed for hydra and mammals where A represents the toxic concentration for adults whereas D represents the toxic concentration affecting development. The hydra assay was sensitive enough to identify among these compounds those capable to pose hazards to development and could be used as predictive of a putative teratogen’s hazard potential (A/D ratio) before performing standard mammal assays (Johnson and Gabel 1983). Later, ‘gastric sections’ have been used as ‘artificial embryo’ by Wilby and Tesh (1990). As the assay based on ‘gastric sections’ was easier to perform than the one using pellets of dissociated cells, it has been reused by several authors (e.g. Pachura-Bouchet 2005; Park and Yeo 2012; Quinn et al. 2008; Vasseur and Pachura 2006). The sensitivity of this ‘hydra artificial embryo’ assay has been previously demonstrated. For example, in the case of nonylphenol exposure, a family of ED compounds found in wastewaters (e.g. Filali-Meknassi et al. 2004), hydra regeneration was disrupted at concentrations three times lower than those leading to adult lethality (Pachura-Bouchet et al. 2006). The present study confirms the sensitivity of this bioassay, and therefore reaffirmed the need of such investigations for risk assessment with environmental concentrations of xenobiotics. The simple diploblastic organization of these animals only consists in two differentiated tissues. Therefore, cnidarians lack complex endocrine system that can be found in vertebrates and several invertebrates (e.g. Arthropoda). Even if physiological regulation and potential disruption are poorly understood in cnidarians, common vertebrate hormones (e.g., steroids, iodinated organic compounds, neuropeptides, and indolamines) have been identified in their tissues and chemical stressors could also impact the physiology of these invertebrates (Tarrant 2005). Keeping in mind that, in contrast to most metazoans, cnidarian cells are not generally organized into organs or systems, with careful consideration, the hydra regeneration assay can be a useful predictor of the potential risk to a developing vertebrate embryo (Bowden et al. 1995). But of course, as pointed out by Tarrant (2005), “care must be taken not to assume processes will be identical in all organisms”. In other words, teratogenic effects on ‘hydra artificial embryo’ exposed to toxicants could reflect teratogenic effect of mammal embryos but no conclusions could be formulated concerning the involved physiological mechanisms. In the present study, in which we have used ‘gastric sections’ as ‘artificial embryo’, teratogenic effects could be observed in several mixtures. The presence of CLD-3Cl had no particular influence on the toxicity of the mixtures, except when both CLD and CLD-3Cl were at their highest levels. For these latter conditions, both compounds could explain the high toxicity of the mixture. Such conditions are however unlikely to occur in the environment as the result of a soil remediation process such as the ISCR one. In first approach, the dechlorination of CLD to generate CLD-3Cl will result in a decrease of CLD concentration in the soil while that of the dechlorination product will increase and so a high concentration of both compounds cannot be present at the same time. In any case, since it will be only possible to apply the remediation processes over a few tens of centimeters of the upper part of the soil, the CLD stock in the lower parts will remain intact. Under such conditions, the work of Ollivier et al (2019, 2020) with column of soils treated in surface by ISCR has clearly shown that the concentrations of CLD-3Cl, but also those of the other dechlorination products will always be lower than CLD in the soil leachates that contaminate the surface and ground waters even if these products are more soluble and so mobile than CLD. Since the ban of CLD field utilization, CLD concentrations in freshwaters could increase solely if CLD stored in soil matrix is released. In a recent study conducted in FWI, Sabatier and colleagues (Sabatier et al. 2021) demonstrate CLD resurgence due to the widespread use of herbicides containing glyphosate since the late 1990s’. This still current agricultural practice is considered to be responsible of a hitherto unseen rise in soil erosion and downstream of a major release of the stable CLD stored in polluted soils since their ban (Sabatier et al. 2021). The severe toxic effects observed at high CLD and CLD-3Cl concentrations on hydra development supports a special warning of agricultural practices that could remobilize CLD and lead to increasing CLD concentrations in freshwaters. In present study, teratogenic effects on ‘hydra artificial embryo’ exposed to the most environmental probable mixtures can be explained by the presence of CLD in the mixture. Indeed, for a given CLD concentration, regeneration scores did not differ when CLD-3Cl concentration increased while scores fluctuate between bad to good when CLD concentrations vary for a given CLD-3Cl concentration. Thus, it seems that the presence of CLD-3Cl in the mixtures, at concentrations expected after the application of a remediation process such as the ISCR, has no supplementary deleterious effect on hydra regeneration capacity. Regeneration scores reflecting no sign of toxicity (score up to 9) can only be observed in the mixtures with CLD concentrations equal or below to 1.02.10−2 µM (5 µg.L−1) (experiments 3, 11 and 14 in figure 2). However, the CLD concentration at 2.88.10−3 µM (1.4 µg.L−1) (experiments 16, 17, 18 in figure 2) appears as a critical concentration leading to decrease of regeneration scores reflecting greater toxic conditions. Interestingly, a CLD concentration which is only 3.5 times higher in the mixture (1.02.10−2 µM, i.e. 5 µg.L−1, experiments 10 and 11 in figure 2) has led to a return to satisfactory regeneration. Hence, our experimental results demonstrate that the toxic effects could not be linked to a progressive gradient of CLD concentration in the mixtures as impairments of regeneration have been observed at both low and high critical concentrations of CLD. The present results clearly demonstrated that the nonmonotonic concentration-response occurred with these organochlorine mixtures because (1) low or high concentrations could lead to the same deleterious damages and (2) because exposition to low and close to low concentrations could lead to good or bad regeneration capacities. Thus, whereas the teratogenic effect could be explained by the presence of CLD in the mixtures, it was not dependent on an increase in CLD concentrations. Such stochastic phenomenon has been also recently described concerning several biological endpoints in hydra entire polyps under exposure to CLD: reproductive rates, morphological changes, and expression of target stress genes (Colpaert et al. 2020). Therefore, our results confirm the previous observations of Colpaert et al. (2020) and lead to the same conclusion: biological effects observed after exposure to CLD follow nonmonotonic dose-response curves. The nonmonotonic dose-response effects are difficult to model, and this phenomenon is a challenge in ecotoxicology because, in absence of a suitable mathematical model to estimate the risk, the presence of such compounds in the environment, whatever their concentrations, represent a threat for exposed populations. Our study pointed out the difficulty to propose a predictive mathematical model when studying compounds belonging to EDs. Without considering the unpredictable stochastic effects of EDs, our experimental design allows to propose an empirical model that can determine the hydra regeneration score (noted Y in the equation of the model) in the domain of interest i.e. for all combinations of concentrations (Figure 6). The approach can also provide information on the most active compounds in the mixtures. From the present data, it can be suggested that our approach is useful to study the mixtures containing compounds having concentration-dependent effects. With a reduced number of experiments, it is possible to model the value of the quantifiable biological responses and to obtain a predictive effect for any combination in the concentration domain. Our method offers an alternative to the isobologram method (Sørensen et al. 2007) as both methods share common advantages such as their simplicity and flexibility (Greco et al. 1995). In a review of studies conducted from 2007 to 2017 either in ecotoxicology or mammalian toxicology (human included), Martin et al. (2021) stated that the isobologram method represents about 12% of mixture investigations. In contrast to our method, the isobologram method needs to determine the biological response beforehand for each compound of the mixture and therefore need more experiments. As our method, the isobologram method could not predict stochastic EDs effects. The isobologram method directly applies the definition of addition of concentrations. Isoboles connect different combinations of doses that produce the same effect and since 2000s’, the effect of the three types of interactions of compounds in mixtures (additivity, antagonism, and synergism) has been determined with the model of Concentration Addition (CA) for substances sharing the same mechanism of action, and under the independent model Action (IA) for those with a different mechanism of action. In our method, there is no need to know the effect of single compound: only the effects of the compounds in mixtures are investigated. Therefore, both methods offer advantages and disadvantages, and the choice of the method depends on the objective of the study. If the objective is to investigate mechanisms of action and determine additive, antagonist or synergistic effects so the isobologram method must be retained. On the other hand, if the objective is to predict the effects of all combinations of mixtures in a range of defined concentrations using a minimum number of experiments, our method is a better choice. The two methods using various mixture ratios are preferable to classical design mixtures. Nevertheless, methods using various mixture ratios represent only 5% of experimental designs (Martin et al. 2021). According to Martin et al. (2021), classical designs with mixtures of two compounds generally consist of a simple combination of concentrations in mixtures. For example: (1) the two chemicals are tested alone at low, medium, and high concentrations, then in mixture, at the same concentrations (e.g. low concentrations together, medium concentrations together and high concentrations together); (2) the concentration of one compound is fixed while the concentrations of the other vary but in a limited number of conditions. A breakthrough in mixture investigations was the possibility to construct response surfaces and this possibility is also offered with our experimental design as well as the isobologram method. However, response surfaces of the biological effects are still under-represented in literature as they represent only 5% of studies from 2007 to 2017 (Martin et al. 2021).
Biological effects of environmental mixtures are puzzling in the cases of EDs and the investigations about EDs challenge the Paracelsus principle “the dose makes the poison”. The European Union regulation states three categories to achieve a good water quality: (1) A river is considered to be in a "good chemical status" if pesticide concentration is not more than 0.1 µg.L-1; (2) a river is considered to be in a "poor chemical status" when the sum of pesticide concentrations exceeds 0.5 µg.L-1; and (3) a river is considered untreatable for water consumption if pesticide concentration exceeds 2 µg.L-1. According to this guideline and when considering only CLD concentrations, all latter categories are encountered in the FWI freshwaters. However, our results indicate that even at low CLD concentrations in the mixtures (0.1 µg.L-1), regeneration capacity of H. vulgaris could be impaired. Previous authors have pointed out that EDs could have higher or similar effects at low concentrations than at higher concentrations (Gore et al. 2015; Vandenberg et al. 2012). In accordance with previous studies using adult hydra polyps (Colpaert et al. 2020), our results show the need to investigate biological effects after exposure to low CLD concentrations. Using the same hydra clone as in present work, Colpaert et al. (2020) have shown asexual reproduction impairments, morphological damages, and modulations of the expression of target genes involved in oxidative stress, detoxification, and neurobiological processes after exposure to low concentrations of CLD (0.1 µg.L-1). In the same way, using the giant freshwater prawn Macrobrachium rosenbergii, a crustacean living in FWI, Gaume et al. (2015) have also pointed out the necessity to investigate biological effects of low environmental concentrations of CLD. Indeed, after exposure to low CLD (0.115 and 0.2 µg.L-1) concentrations, these authors have demonstrated an induction of genes involved in biotransformation processes and in defense mechanisms against oxidative stress in M. rosenbergii (Gaume et al. 2015). Thus, in agreement with previous studies, our results confirm the need for a better knowledge of the stochastic effects of EDs.