Importance of Toxicokinetics for Interspecies Variation in Sensitivity to

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The importance of toxicokinetics for interspecies variation in sensitivity to chemicals Anna-Maija Nyman, Kristin Schirmer, and Roman Ashauer Environ. Sci. Technol., Just Accepted Manuscript • DOI: 10.1021/es5005126 • Publication Date (Web): 23 Apr 2014 Downloaded from http://pubs.acs.org on May 1, 2014

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Environmental Science & Technology

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The importance of toxicokinetics for interspecies variation in sensitivity to chemicals

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Anna-Maija Nyman †, ‡, §, ||, *, Kristin Schirmer †, ||, #, Roman Ashauer †, ǂ

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†(a)

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Science and Technology, 8600 Dübendorf, Switzerland

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‡

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Research, 04318 Leipzig, Germany

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§

Department of Environmental Toxicology, Eawag - Swiss Federal Institute of Aquatic

Department of Ecological Modelling, UFZ - Helmholtz Centre for Environmental

Department of Biology, University of Eastern Finland, 80101 Joensuu, Finland

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||(a)

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Switzerland

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#

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Lausanne, Switzerland

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ǂ

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* Corresponding author: [email protected]

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(a) places where work was performed

Institute of Biogeochemistry and Pollutant Dynamics, ETH Zürich, 8092 Zürich,

School of Architecture, Civil and Environmental Engineering, EPF Lausanne, 1015

Environment Department, University of York, Heslington, York, UK

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KEYWORDS: aquatic invertebrates, internal distribution, quantitative whole body

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autoradiography, species traits, pesticides


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ABSTRACT

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Interspecies variation in sensitivity to synthetic chemicals can be orders of magnitude large. Species

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traits causing the variation can be related to toxicokinetics (uptake, distribution, biotransformation,

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elimination), or toxicodynamics (interaction with biological target sites). We present an approach to

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systematically measure and model the contribution of uptake, biotransformation, internal

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distribution and elimination kinetics towards species sensitivity differences. The aim is to express

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sensitivity as target tissue specific, internal lethal concentrations. A case study with the pesticides

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diazinon, imidacloprid and propiconazole and the aquatic invertebrates Gammarus pulex,

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Gammarus fossarum and Lymnaea stagnalis illustrates the approach. L. stagnalis accumulates more

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pesticides than Gammaridae when measured in whole organisms but less in target tissues such as

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the nervous system. Toxicokinetics, i.e. biotransformation and distribution, explain the higher

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tolerance of L. stagnalis to the insecticide diazinon when compared to Gammaridae. L. stagnalis

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was again more tolerant to the other neurotoxicant imidacloprid, however, the difference in

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sensitivity could not be explained by toxicokinetics alone, indicating the importance of

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toxicodynamic differences. Sensitivity to propiconazole was comparable among all species, and

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when expressed as internal lethal concentrations, falls in the range of baseline toxicity.

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INTRODUCTION

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Almost 7000 aquatic invertebrate species live in European waters and they play an important role in

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nutrient cycles, primary production, decomposition and translocation of materials1, 2. Invertebrates

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are facing a major challenge due to exposure to various pesticides which are entering surface waters

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after drift, leaching or run-off from fields3, 4. Species do not respond to pollution similarly; rather a

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large variation in sensitivity among organisms exists. Not only do species vary substantially in their

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sensitivities to a given toxicant, but a given species can vary tremendously in its sensitivity across

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toxicants5-9. Interspecies variation in sensitivity is reported to be lower for baseline toxicants, which

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act by disturbance of membrane intergrity and functioning, than for those which exhibit a more

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specific mode of action5,

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based on their hydrophobicity or partitioning to membrane vesicles. Reactive compounds and

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compounds interacting specifically with certain receptors are more toxic and the more specifically

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toxic they are, the more variation in sensitivity is observed11.

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. Therefore the toxicity of baseline toxicants can be well predicted

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The differences in sensitivity among species can be related to toxicodynamics, which

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comprise the processes taking place at the target sites creating the toxic effects and its propagation

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to whole organism responses13. Species traits related to toxicodynamics, such as presence or

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absence of target receptors, are important in causing interspecies variation in sensitivity to

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toxicants11, 12. However, many studies have also highlighted the importance of toxicokinetics10, 12, 14-

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determine the chemical concentrations which reach the target sites13. For example, the tissue residue

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approach is based on linking the effects to the amount of chemical present in the organism or

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specifically at the sites of toxic action. This approach can substantially reduce the apparent

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variability among species by bringing the metric of effective dose closer to the interaction of the

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toxicant and the target receptor17.

, which comprise chemical uptake, biotransformation, distribution and elimination and therefore


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The differences in toxicokinetics have been shown to occur as differences in uptake and

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elimination rates6,

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respiratory strategy14, 16, 18, 22. In addition, the biotransformation capacity of a species to inactivate or

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activate specifically acting compounds has been considered an important factor causing differences

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in sensitivity12, 23. Differences in toxicokinetics have been shown to explain some of the variation in

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sensitivity (50-60%18) but establishing the link from differences in toxicokinetics to interspecies

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variation in sensitivity has been partly hindered by a lack of important toxicokinetic information.

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First, biotransformation rates of the parent compound to more toxic metabolites can differ widely

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among species. However, the biotransformation rates have not been measured in all studies10, 18.

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Second, earlier studies have been focusing on comparing the sensitivities with the total body

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burden, assuming that concentration in the whole organism describes or correlates with the

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concentrations at the target sites16, 18. Whole body based toxicokinetic data can be misleading. For

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example, a compound may preferentially accumulate in non-target tissues and therefore the

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concentrations extracted from whole organisms do not necessarily reflect the concentrations at the

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target sites.

which arise from the differences in e.g. lipid content, body size and

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This study investigates the differences in sensitivity to three pesticides among three aquatic

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invertebrates. We selected two neurotoxic insecticides diazinon and imidacloprid, and the fungicide

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propiconazole. The objectives were to (i) explore how much the consideration of tissue residues can

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reduce interspecies variation in sensitivity by systematically eliminating toxicokinetic based

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variability arising from differences in chemical bioaccumulation in whole organisms (uptake,

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elimination), biotransformation, and distribution into the target tissues (ii) quantify the differences

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which remain in toxicodynamics when detailed toxicokinetics have been considered.

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The selection of species covered two closely related species Gammarus pulex and

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Gammarus fossarum, which are similar in their physiology and likely have similar target receptors,

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while the third selected species was the phylogenetically distant species Lymnaea stagnalis, which 5 ACS Paragon Plus Environment


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differs by its physiology (e.g. breathing mechanism) as well as possibly by its target receptors.

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Therefore our species selection aimed to provide insight into how the processes causing variation in

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sensitivity differ between phylogenetically close to distant species. We use not only already

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established methods to study chemical bioaccumulation and biotransformation but also apply novel

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methodology to investigate chemical distribution to target tissues in invertebrates.

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MATERIAL AND METHODS

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Chemicals

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We used mixtures of

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from Sigma-Aldrich (Buchs, Switzerland) and radiolabeled chemicals from the Institute of Isotopes

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Co., Ltd. (Budapest, Hungary). Radiochemical purities were above 99.67% (propiconazole),

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96.97% (imidacloprid) and 99.21% (diazinon). Unlabeled chemical purities were above 98.4%

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(propiconazole), 99.9% (imidacloprid) and 98.2% (diazinon).

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C-labeled and unlabeled chemicals. Unlabeled chemicals were purchased

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Both, diazinon (log Kow 3.69) and imidacloprid (log Kow 0.57) are neurotoxicants but act via

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different mechanisms. Diazinon is an acetylcholinesterase (AChE) inhibitor while imidacloprid

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interferes with insect nerve impulses by binding to nicotinic acetylcholine receptors in the central

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nervous system24. In order to inhibit AChE efficiently, diazinon requires metabolic transformation

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to its oxon derivative, diazoxon, because the P=O group binds more tightly to AChE than the P=S

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group of the parent diazinon24, 25. Due to inhibition of AChE, acetylcholin is not degraded from the

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synapsis and the persistent activation leads to overstimulation by the neurotransmitter. Similar

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effects are observed under exposure to imidacloprid as the compound cannot be hydrolyzed by ACh

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esterase from the nicotinic ACh receptors26, 27. Propiconazole (log Kow 3.72) acts by inhibiting the

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enzyme sterol 14α-demethylase which belongs to the superfamily of cytochrome P450 enzymes28.

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In animals, sterol 14α-demethylase action is part of the pathway that leads to cholesterol 6 ACS Paragon Plus Environment

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biosynthesis. Nevertheless, under short-term exposure, propiconazole is hypothesized to act via

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narcosis, at least in Gammarids29.

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Species

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We investigated three aquatic invertebrate species: Gammarus pulex, Gammarus fossarum and

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Lymnaea stagnalis. Gammarus pulex and G. fossarum (Crustacea, Amphipoda, Gammaridae) are

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related on the genus level but the freshwater gastropod L. stagnalis (Gastropoda, Pulmonata,

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Basommatophora) represents a different phylum amongst animals. All three species are abundant in

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European freshwaters and play an important role in the decomposition of organic material and serve

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as a food source for organisms of higher trophic levels1. We used subtype A of Gammarus fossarum

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in our experiments; the subtype was determined by pyrosequencing30. The Gammaridae test

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individuals in our study were collected from small headwater streams in Switzerland (G. pulex:

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Itziker Ried, 47.27463 N, 8.78923 E; G. fossarum: Kollbrunn, 47.46361 N, 8.80048 E). The

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freshwater snail L. stagnalis was cultured in the laboratory and the population was initiated from

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egg clutches of the strain reared in the INRA Experimental Unit of Aquatic Ecology and

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Ecotoxicology (Rennes, France) which originates from natural ponds in Le Rheu, France31. The

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experiments on Gammaridae species were conducted at +13ºC (12:12 light:dark) and on Lymnaea

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at +20ºC (14:10 light:dark). For food and shelter, we provided uncontaminated horse chest-nut

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(Aesculus hippocastanum) leaves inoculated with the fungi Cladosporium herbarum for

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Gammaridae and lettuce (Lactuca sativa) for Lymnaea ad libitum. More information on the culture

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and handling of the test organisms is provided in the Supporting Information (SI).

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Metabolite screening



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The susceptibility of the pesticides to biotransformation in our test species was determined for all

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compounds by carrying out metabolite screening experiments. The organisms were exposed to the

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pesticides for 24 hours, the organisms were sampled, and the samples were frozen until analysis.

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The analysis was conducted according to our previous studies using high-performance liquid

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chromatography (HPLC) connected with a radiodetector29, 32. For snail samples, the shell length

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was measured and the organism was separated from the shell before the extraction. See more details

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in the SI.

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Bioaccumulation based on total body burden

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We conducted bioaccumulation experiments only with Gammarus fossarum and Lymnaea

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stagnalis, as bioaccumulation of diazinon, imidacloprid and propiconazole in Gammarus pulex have

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been published elsewhere29, 33, 34. The experiments included a 1 day exposure to a pesticide followed

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by a depuration period up to 7 days, with the duration depending on the species and chemical of

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concern. In all toxicokinetic experiments (total bioaccumulation, chemical distribution) we used

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concentrations which were well below one day acute toxicity based on our toxicity tests but still

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high enough to measure the internalized chemicals. During both the uptake and depuration phases,

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test media and organisms were sampled for determination of chemical concentrations at different

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time points. If metabolites were observed in the metabolite screening experiment, the animals

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sampled during toxicokinetic experiments were analysed for both, parent pesticide and its

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metabolite concentrations using HPLC as in metabolite screening experiments. If no

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biotransformation products were observed, only the total body burden was measured, similarly to

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the method described by Ashauer and co-workers using liquid scintillation counting33. See SI for

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description of the analyses.

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The differences in uptake and elimination of the compounds amongst our test species was

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investigated by calibrating toxicokinetic models using the data from bioaccumulation experiments


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(i.e. internal and external concentrations over time). We assume that all measured radioactivity

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contributes to whole body concentrations. Thus for Gammaridae we include chemicals adsorbed to

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the body surface, whereas in the case of snails we removed the shells before extraction and

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therefore did not include chemicals adsorbed to the shell. In case of propiconazole and

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imidacloprid, a one-compartment toxicokinetic model without separation of the parent pesticide and

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its biotransformation products (Eqn 1) was calibrated using total body burden obtained from liquid

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scintillation counting33. For imidacloprid this model was chosen because the compound was not

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biotransformed in any of the test organisms (Figs S1-S2 and

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was biotransformed in Gammaridae but not in L. stagnalis (Figs S3-S4 and

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study indicated that propiconazole acts as a baseline toxicant in G. pulex and, since we do not have

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further knowledge about the toxic mechanism of the biotransformation products in Gammaridae, we

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assumed that the biotransformation products also act as baseline toxicants. Therefore we used the

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one-compartment toxicokinetic model also for propiconazole.

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). Propiconazole on the other hand

dC int (t ) = C ext (t ) ⋅ k in − C int (t ) ⋅ k out dt

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). As our previous

Eqn 1

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where Cint (t) is the internal chemical concentration in the organisms [nmol/kgwet weight], Cext (t) is the

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concentration in water [nmol/L], kin is the uptake rate constant [L·kg-1·d-1], kout is the elimination

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rate constant [1/d], and t is time [d].

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In the case of diazinon, diazoxon (2-isopropyl-6-methyl-4-pyrimidinol) and the inactive

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metabolite pyrimidinol (2-isopropyl-4-methyl-6-hydroxypyrimidine) were measured in all our test

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species (Figs S5-S6 and

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this biotansformation product to our toxicokinetic model and describe the kinetics with a two

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compartment model (Eqns 2-3).

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d C int diazinon (t ) = C ext (t ) ⋅ k in − C int diazinon (t ) ⋅ k out diazinon − C int diazinon (t ) ⋅ k activation dt

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). As the active form of diazinon is the metabolite diazoxon, we added

Eqn 2 9

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dCint diazoxon (t ) = C int diazinon (t ) ⋅ k activation − Cint diazoxon (t ) ⋅ k out diazoxon dt

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Eqn 3

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where Cint diazinon (t) is the internal diazinon concentration in organisms [nmol/kgwet weight], Cint diazoxon

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(t) is the internal diazoxon concentration in organisms [nmol/kgwet

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concentration in water [nmol/L], kin is the uptake rate constant [L·kg-1·d-1], kout diazinon is the diazinon

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elimination rate constant [1/d], kactivation is the activation rate of diazinon to diazoxon [1/d], kout

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diazoxon

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can be eliminated from organisms via a variety of mechanisms35. Therefore the diazinon and

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diazoxon elimination rate constants were calibrated to the disappearance of the compounds from the

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organisms and as such also represent the loss via biotransformation to e.g. pyrimidinol in addition

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to depuration.

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weight],

Cext (t) is the diazinon

is the diazoxon elimination rate constant [1/d], and t is time [d]. Diazinon and its metabolites

Bioaccumulation data, fitting of the toxicokinetic model and a summary of the goodness of fit of all calibrations are illustrated in the SI.

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The toxicokinetic parameters were used to estimate bioaccumulation factors, BAFs [L/kg], for

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the whole organism, which represent the bioaccumulation in steady state conditions. The BAFs for

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imidacloprid and propiconazole were calculated as:

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=

Eqn 4

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For diazinon, BAFs were calculated as:

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=

Eqn 5

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The formation and retention of the metabolite diazoxon was calculated as retention potential

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factor (RPF):

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=

Eqn 6

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In steady-state the RPFdiazoxon equals the diazoxon concentration in the organism divided by

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the diazinon concentration in the organism. The BAF and RPF were further used to calculate MEF,

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the metabolite enrichment factor. This factor quantifies the ratio of the metabolite concentration in

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the organism over the concentration of the parent compound in the medium (MEF = BAF × RPF),

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not only the formation of metabolites from parent compound as in case of the RPF32.

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Chemical distribution

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To determine the internal distribution of the compounds, we analyzed chemical concentrations in

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organisms using Quantitative Whole Body Autoradiography (QWBA)36. Organisms were exposed

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to each of the pesticides for 24 hours. Five replicates of each species (Lymnaea stagnalis,

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Gammarus pulex, Gammarus fossarum) containing a 14C-labeled pesticide and its metabolites were

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frozen in -80 ºC until analysis, which was performed at Harlan Laboratories Ltd. Itingen,

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Switzerland. The specimens were embedded in a section frame and cryo-sections of 40 (Lymnaea,

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Frame 1) or 20 (Gammaridae, Frame 2) µm thickness at different levels of the body were produced.

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Gammaridae individuals are smaller than Lymnaea and therefore, in order to capture all possible

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organs in the animals, the section thickness was decreased. Concentrations of radioactivity in

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tissues and organs were determined in 12 horizontal (Lymnaea) or 20 vertical (Gammaridae)

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sections using the radioluminography technique (RLG). The sections were exposed to an imaging

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plate and the tissues and organs of interest were marked as area on the radioluminographs and the

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concentration of radioactivity was calculated as integral of the marked area expressed in dpm/mm².



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The concentration of radioactivity in tissues and organs was calculated in mol/g, based on

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calibration curve and the thickness of the sections. See more details in the SI.

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Because the exposure concentration varied between different species and chemicals, we

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present the chemical distribution data as bioaccumulation factors, BAFs [L/kg], which is calculated

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by dividing internal concentrations [µmol/kg] by external concentrations [µmol/L]. Since the

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QWBA technique measures only the total radioactivity, for diazinon we calculated also the amount

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of diazoxon from total radioactivity by multiplying the ratio of diazoxon/total diazinon with the

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concentrations in different tissues. The ratio of diazoxon/total diazinon was obtained from a TK

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model (see description above), calibrated using whole body bioaccumulation data, simulating

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internal concentrations of diazinon and diazoxon in the exposure scenario used in the QWBA

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experiment for each species (24-h exposure). Thus we were forced to assume that the ratio of

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diazinon to diazoxon does not vary amongst different tissues. This assumption might cause

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uncertainty in our results because cytochrome P450 activity, which is responsible for

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biotransforming diazinon to diazoxon, has been shown to be higher in digestive organs, such as

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digestive glands in gastropods or hepatopancreas in crustaceans, than in other tissues37. Even

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though the P450 enzymatic levels in these digestive organs are comparable in gastropods and

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crustaceans, there are differences in lipid content of the organs among these two species classes

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(higher in digestive glands of gastropods than in hepatopancreas of crustaceans38, 39). This in turn

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might affect further distribution of biotransformation products from the digestive tissues throughout

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the organism and cause uncertainty when we compare diazoxon concentrations among our test

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species in different organs using the constant ratio of diazoxon/diazinon. However, the P450

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pathway occurs also in other tissues37 and our approach is the best approximation of diazoxon

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distribution in tissues that we can achieve without direct tissue specific quantification of

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metabolites.

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Toxicity experiments

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The acute toxicity experiments consisted of five to seven pesticide concentrations with two to four

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replicate beakers each. Pesticide concentrations in water were measured and the survival of

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organisms was observed daily for four days (see SI). Note that steady state was not possibly reached

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in this time course. From the data we calculated the lethal concentrations for 50% of the tested

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populations (LC50 values) (sigmoidal dose response curve, variable slope, GraphPad Prism 4.03,

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GraphPad Software Inc., San Diego, USA).

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Internal lethal concentrations

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To determine the degree to which interspecies variation in sensitivity is caused by toxicokinetics,

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we estimated internal lethal concentrations (ILC50 values13, 40) for all combinations of species and

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chemicals. The ILC50 values were calculated as product of the lethal exposure concentration

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(LC50) and the bioaccumulation factor (BAF or MEF for diazoxon), using both total body burden

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and target tissue specific bioaccumulation. In case of diazinon and imidacloprid, the target tissue is

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known to be specific (i.e. nervous system); however, propiconazole likely acts as a baseline toxicant

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under short term exposures and the target sites are cell membranes in all tissues29.

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RESULTS AND DISCUSSION

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Bioaccumulation based on total body burden

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Toxicokinetic processes such as biotransformation and uptake of chemicals vary among

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invertebrates19-21. Already among arthropods, bioconcentration of chlorpyrifos varied from 100 to

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almost 14,000 L/kg wet weight18. However, bioaccumulation is reported to be similar among



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related, compared to unrelated species, and especially the elimination rates among related species

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appear to be similar15. This was supported by our study: bioaccumulation of the pesticides was

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almost identical for Gammaridae (Table 1).

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Regarding the unrelated species, we observed that all compounds accumulate in Lymnaea

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more than in Gammaridae (Table 1). In case of propiconazole (log Kow 3.72), Lymnaea had both, a

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larger uptake rate and also a smaller elimination rate which results in larger bioaccumulation.

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Diazinon was biotransformed to diazoxon and pyrimidinol in all species (Figs S5-S6). Based on

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total body burden, the metabolite enrichment factor, MEF, of diazoxon showed that L. stagnalis

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accumulated this toxic diazinon metabolite the most from all species investigated (Table 1).

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Imidacloprid (log Kow 0.57) is taken up rapidly by Lymnaea which resulted in a large

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bioaccumulation factor when compared to Gammaridae (Table 1); however, due to its large

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elimination rate constant, the compound is eliminated from Lymnaea faster than from Gammaridae

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after exposure is terminated (Figs S10-S11). One possible explanation for the fast uptake and

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elimination from Lymnaea is that we fed the animals during the experiments, which means

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chemicals attached to food influence our results.

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Internal distribution

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Total body burden does not necessarily reflect the concentrations at the target sites due to

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accumulation of the chemical in non-target tissues, such as the exoskeleton or adipose tissue. To

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link the lethal body burden to target site specific chemical concentrations, we investigated chemical

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distribution using Quantitative Whole Body Autoradiography (QWBA). As seen in Figure 1, the

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highest concentrations of the chemicals can be found in the gastrointestinal complex of Lymnaea

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while in Gammaridae the chemicals are more evenly distributed. Indeed, with the exception of the

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gut/gastrointestinal complex, the tissue specific BAFs for Lymnaea are smaller than for


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Gammaridae (Figure 2). Breathing mechanism as a biological trait has been suggested to cause the

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differences in bioaccumulation and toxicity of pollutants among species14, 16, 20, 22, and can explain

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the low accumulation in tissues of the snail observed in this study. The snail breathes mainly air41

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while Gammaridae take up oxygen from water via gills42 and thus Gammaridae acquire the

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chemicals via their gills in addition to uptake via food. An alternative explanation may be related to

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differences in lipid content of different organs. The digestive glands (e.g. hepatopancreas) of

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Lymnaea contain large amounts of lipids38, which might limit the distribution of chemicals from the

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digestive tract to other tissues. This is supported by the higher BAFs for Lymnaea in the

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gut/gastrointestinal complex when compared to Gammaridae - however, we cannot distinguish how

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much of the activity is in the food and which fraction is truly internalized.

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The QWBA method has been established for rats and other vertebrates36 which are easier to

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handle during embedding and the organs are more visible than in invertebrates. Due to the small

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size of our test organisms, some of the organs could not be detected in all samples and the number

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of replicate samples per organ type were sometimes very small (see SI). Another uncertainty related

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to the results of our distribution analysis is caused by the duration of our experiment. All species

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were exposed to the chemicals only for one day and therefore the steady state was not reached in

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some of them (Table S2). Essentially our QWBA provides only one snapshot into a dynamic

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system. Due to these two factors, small sample size and duration of exposure, the results for

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chemical distribution should be treated with some caution. However, we see patterns in the

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distribution data which show that the method is very informative. For instance, all compounds

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showed smaller accumulation to tissues of Lymnaea (Figure 2), even though Lymnaea shows higher

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bioaccumulation in steady state as well as after 24 hours based on our results on total body burden

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(Table S2). Being very informative, we encourage further research to use QWBA methods to take

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into account chemical distribution in invertebrates. However, we recommend increasing the sample



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size in order to find all relevant tissues in sufficient replication, as well as sampling at different

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times until reaching steady state.

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Toxicity

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A large number of ecotoxicological studies has shown both interspecies and interchemical variation

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in response to chemical exposure (e.g.

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toxicants (i.e. membrane perturbation) is very small8, 11, 12, in general less than a factor of ten5. We

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performed acute toxicity experiments and showed that the acute toxicity of the fungicide

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propiconazole is comparable in all species investigated here, indeed within a factor of 10 (Figure 3,

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left column), which indicates that propiconazole likely acts via narcosis under short-term exposure,

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as already suggested in our earlier study with G. pulex29.

6-9

). The interspecies variation in response to baseline

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Diazinon and imidacloprid on the other hand are specifically acting insecticides, both

336

affecting the nervous system. Diazinon is an AChE-inhibiting organophosphate and the interspecies

337

variation to diazinon exposure is very large11. Imidacloprid affects the nervous system in a different

338

way, by binding to cholinergic receptors, and the toxicity can vary by one order of magnitude

339

already between species belonging to the same class (Arthropoda, Crustacea, Malacostraca)43. We

340

showed here that Lymnaea stagnalis is more tolerant to diazinon and imidacloprid than

341

Gammaridae species (Figure 3, left column). This finding is in good agreement with the fact that

342

these compounds are insecticides and Gammarus pulex and Gammarus fossarum (Arthropoda,

343

Crustacea, Amphipoda, Gammaridae), as belonging to the phylum Arthropoda, are more affected

344

than a mollusc, like Lymnaea stagnalis (Mollusca, Gastropoda, Pulmonata, Basommatophora).

345

Accordingly, sensitivity to toxicants has been suggested to be driven by phylogenetically linked

346

physiological traits such as respiratory strategy, excretion mechanisms and presence of

347

biotransformation enzymes and pathways6, 15, 21, as well as target site structure44. These hypotheses 16 ACS Paragon Plus Environment

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348

are supported by our study since related Gammaridae species respond to chemical stress similarly.

349

To better understand which traits are governing the differences in sensitivity among our test species,

350

we combined the information on chemical bioaccumulation and distribution with the toxicity data.

351

352

Internal lethal concentrations

353

To account for the differences in bioconcentration when comparing the sensitivity of our

354

invertebrates, we calculated internal LC50 values (ILC50 = bioaccumulation factor BAF x LC50)

355

using whole body based and tissue specific BAFs (MEFs for diazoxon). Below we compare the

356

species sensitivities separately to each compound using this approach.

357

When we calculate internal LC50 concentrations for propiconazole based on the BAF for the

358

whole body, the ILC50 values differ more between species than LC50 values alone (Figure 3, lower

359

panel, middle). However, chemicals can stay in the gut where they are not truly internalized (Figure

360

2) or accumulate to other tissues which are not relevant considering the effects, such as adipose

361

tissue or exoskeleton. Therefore the ILC50 values must be calculated based on tissue specific data

362

instead of the whole body BAFs. Target sites for propiconazole are cell membranes present in all

363

tissues, and therefore using any generic tissue type, such as muscle which we used, to compare

364

ILC50 values is better than using whole body residues. When based on muscle tissue BAFs, the

365

ILC50 values for propiconazole are again very similar amongst species (Figure 3, lower panel,

366

right). Baseline toxicants have small interspecies variation in sensitivity5,

367

species responded similarly to propiconazole, the fungicide likely acts via narcosis in acute lethal

368

exposure. This was also supported by the comparison between the ILC50 values of our test species

369

and ILC range for known baseline toxicants in Daphnia magna13,

370

Figure 3).

40

11, 12

, and as our test

(solid and dashed lines in



Page 18 of 29

371

Nevertheless, even though propiconazole acts as a baseline toxicant in our short-term

372

experiments, we cannot exclude the possibility of specific effects of the chemical under long term

373

exposures. Propiconazole inhibits cytochrome P450 enzymes, which are responsible for

374

metabolizing a wide range of endogenous and xenobiotic compounds45. Therefore the inhibition of

375

these enzymes can lead, for example, to decreased biotransformation capability which, in chemical

376

mixtures, can enhance the toxicity of those chemicals normally detoxified by the P450 system. The

377

gastrointestinal complex and hepatopancreas of L. stagnalis and gut and hepatopancreas of

378

Gammaridae are major digestive organs with high cytochrome P450 activity37, 45. As we observed

379

high bioaccumulation of propiconazole in these organs (Table S4), the risk of propiconazole in long

380

term exposure might be considerable.

381

When calculating the ILC50 total body values for the insecticide imidacloprid, we see that the

382

differences in sensitivity among unrelated species is as large as based on external concentrations

383

(Figure 3, middle panel, left). However, when we correct the bioaccumulation to target site specific

384

BAFs (i.e. nervous system BAFs), the differences become smaller, especially when we compare

385

between the two Gammaridae. Thus, in this case, bioaccumulation to target sites explains the

386

differences among Gammaridae but it can not explain alone why Lymnaea is so insensitive towards

387

imidacloprid. Additional factors, such as differences in the nervous system likely play a role here. It

388

has been shown that multiple binding subunits of imidacloprid exist in the nicotinic acetylcholine

389

receptors and some species might have more high-affinity subunits than others46. For instance, the

390

order Hemiptera (where the target organisms belong) is suggested to have more of the high-affinity

391

binding sites than nonhemipteran species. The subunits of nicotinic ACh receptors existed already

392

in early ancestors (Bilateria) but started to diverge since then44. As Gammaridae belong to the same

393

phylum (Arthropoda) as hemipteran species, the composition of the nicotinic receptor subunits is

394

likely more similar between Gammaridae and Hemiptera than between the mollusk Lymnaea and

395

hemipterans. Accordingly, we observed higher sensitivity for Gammaridae than Lymnaea. The 18 ACS Paragon Plus Environment

Page 19 of 29


396

variation of the target site based ILC50 values among Gammaridae are in the same range as the

397

variation in case of the baseline toxicant propiconazole, suggesting that these related species have

398

similar amounts of high affinity nicotinic receptor subunits.

399

When comparing the sensitivity between G. pulex and G. fossarum to diazinon,

400

bioaccumulation in the whole body can explain differences in sensitivity: G. pulex, which is more

401

sensitive to diazinon (Figure 3, top panel, left), accumulates more of the toxic metabolite diazoxon

402

than G. fossarum (MEF, Table 1). When the toxicokinetic parameters (Table 1) are compared, we

403

see that G. pulex possibly accumulates more because of the slower elimination of diazoxon. On the

404

other hand, as with imidacloprid, whole-body toxicokinetics cannot explain why Lymnaea is much

405

more tolerant to diazinon than Gammaridae (Figure 3, top panel, left). Rather the opposite, the

406

mollusk accumulates diazoxon in the whole body more than Gammaridae (MEF, Table 1). The

407

explanation, however, can be found in the chemical distribution within organisms because the

408

bioaccumulation in nervous tissue is much less in Lymnaea than in Gammaridae (Figure 2). Using

409

the chemical distribution data and MEFs in nervous tissue to calculate target tissue specific ILC50

410

values, we see that the response to diazinon among our test species becomes much more similar

411

when compared to external LC50 or ILC50 total body values (Figure 3, top panel). In fact, the

412

variation in ILC50 values among species becomes similar to that of the baseline toxicant

413

propiconazole (SE values, Figure 3). Therefore we can conclude that differences in toxicokinetics

414

cause the variation in sensitivity to diazinon even amongst the unrelated invertebrates studied here,

415

a mollusk and arthropods. However, even though we could bring the interspecies variation in the

416

sensitivity to diazinon to the levels of a baseline toxicant by using detailed toxicokinetic data, we

417

cannot exclude the possibility that the remaining variation may be due to differences in

418

toxicodynamics, e.g. whole-body background AChE activity11.

419



Page 20 of 29

420

Application and implications

421

We presented an approach to systematically measure and model the contribution of uptake,

422

biotransformation, internal distribution and elimination kinetics towards species sensitivity

423

differences. Our approach can help to better distinguish baseline toxicants from specifically acting

424

compounds and identify causes of species sensitivity differences.

425 426

Acknowledgements We thank Virginie Ducrot and INRA for providing egg clutches of Lymnaea

427

stagnalis to establish the culture at Eawag. We are also grateful for our colleagues in the department

428

of Aquatic Ecology at Eawag for providing Lymnaea test media, helping with determination of the

429

tissues in snails and Gammaridae and for determination of the subtype of Gammarus fossarum.

430

Many thanks to colleagues at the Environmental Toxicology department at Eawag for their help in

431

Gammarus hunting and maintaining the Lymnaea culture. Harlan Laboratories, Stephan Hassler and

432

Anke Henninger are acknowledged for the cooperation in developing and applying the QWBA

433

technique to invertebrates. This research has been financially supported by the European Union

434

under the 7th Framework Programme (project acronym CREAM, contract number PITN-GA-2009-

435

238148).

436

437

Supporting Information Available: Supporting information includes more experimental details,

438

radio-HPLC chromatograms, plots of toxicokinetic model fits and data, measured external and

439

internal concentrations over time, and detailed organ specific BAFs. This information is available

440

free of charge via the Internet at http://pubs.acs.org/.


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441

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19. Jeon, J.; Kurth, D.; Hollender, J., Biotransformation pathways of biocides and pharmaceuticals infreshwater crustaceans based on structure elucidation of metabolites using high resolution mass spectrometry. Chemical Research in Toxicology 2013, 26, (3), 313-324. 20. Meredith-Williams, M.; Carter, L. J.; Fussell, R.; Raffaelli, D.; Ashauer, R.; Boxall, A. B. A., Uptake and depuration of pharmaceuticals in aquatic invertebrates. Environmental Pollution 2012, 165, (0), 250258. 21. Livingstone, D. R., The fate of organic xenobiotics in aquatic ecosystems: quantitative and qualitative differences in biotransformation by invertebrates and fish. Comparative Biochemistry and Physiology Part A: Molecular & Integrative Physiology 1998, 120, (1), 43-49. 22. Buchwalter, D. B.; Jenkins, J. J.; Curtis, L. R., Respiratory strategy is a major determinant of [3H]water and [14C]chlorpyrifos uptake in aquatic insects. Canadian Journal of Fisheries and Aquatic Sciences 2002, 59, (8), 1315-1322. 23. Chambers, J. E.; Carr, R. L., Biochemical mechanisms contributing to species differences in insecticidal toxicity. Toxicology 1995, 105, (2–3), 291-304. 24. Yu, S., The toxicology & biochemistry of insecticides Taylor & Francis Group, LLC, USA 2008, 296. 25. Ankley, G. T.; Dierkes, J. R.; Jensen, D. A.; Peterson, G. S., Piperonyl butoxide as a tool in aquatic toxicological research with organophosphate insecticides. Ecotoxicology and Environmental Safety 1991, 21, (3), 266-274. 26. Jeschke, P.; Nauen, R., Nervous System. In Modern Crop Protection Compounds, Krämer, W.; Schirmer, U., Eds. Wiley-VCH Verlag GmbH: Wenheim, Germany, 2008; pp 927-1088. 27. Abbink, J., The biochemistry of imidacloprid. Pflanzenschutz-Nachrichten Bayer 1991, 42, 183-195. 28. Zarn, J. A.; Brüschweiler, B. J.; Schlatter, J. R., Azole fungicides affect mammalian steroidogenesis by inhibiting sterol 14Î±-demethylase and aromatase. Environmental Health Perspectives 2003, 111, (3), 255-261. 29. Nyman, A.-M.; Schirmer, K.; Ashauer, R., Toxicokinetic-toxicodynamic modelling of survival of Gammarus pulex in multiple pulse exposures to propiconazole: model assumptions, calibration data requirements and predictive power. Ecotoxicology 2012, 21, (7), 1828-1840. 30. Westram, A. M.; Jokela, J.; Baumgartner, C.; Keller, I., Spatial distribution of cryptic species diversity in european freshwater amphipods Gammarus fossarum as revealed by pyrosequencing. PLoS ONE 2011, 6, (8), e23879. 31. Coutellec, M.-A.; Lagadic, L., Effects of self-fertilization, environmental stress and exposure to xenobiotics on fitness-related traits of the freshwater snail Lymnaea stagnalis. Ecotoxicology 2006, 15, (2), 199-213. 32. Ashauer, R.; Hintermeister, A.; O'Connor, I.; Elumelu, M.; Hollender, J.; Escher, B. I., Significance of xenobiotic metabolism for bioaccumulation kinetics of organic chemicals in Gammarus pulex. Environmental Science & Technology 2012, 46, 3498-3508. 33. Ashauer, R.; Caravatti, I.; Hintermeister, A.; Escher, B. I., Bioaccumulation kinetics of organic xenobiotic pollutants in the freshwater invertebrate Gammarus pulex modeled with prediction intervals. Environmental Toxicology and Chemistry 2010, 29, (7), 1625-1636. 34. Ashauer, R.; Hintermeister, A.; Caravatti, I.; Kretschmann, A.; Escher, B. I., Toxicokinetic and toxicodynamic modeling explains carry-over toxicity from exposure to diazinon by slow organism recovery. Environmental Science & Technology 2010, 44, (10), 3963-3971. 35. Larkin, D.; Tjeerdema, R., Fate and effects of diazinon. Rev Environ Contam Toxicol 2000, 166, 49-82. 36. Coe, R. A. J., Quantitative Whole-Body Autoradiography. Regulatory Toxicology and Pharmacology 2000, 31, (2), S1-S3. 37. Solé, M.; Livingstone, D. R., Components of the cytochrome P450-dependent monooxygenase system and ‘NADPH-independent benzo[a]pyrene hydroxylase’ activity in a wide range of marine invertebrate species. Comparative Biochemistry and Physiology Part C: Toxicology & Pharmacology 2005, 141, (1), 20-31. 38. Nagare, K. R.; Dummalod, C. B., Alterations in lipid contents of infected and non-infected snail Lymnaea acuminata during patency period from Aurangabad (M.S.). Indian Journal of Applied Research 2013, 3, (4), 14-15.


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541 542 543 544 545 546 547 548 549 550 551 552 553 554 555 556 557 558 559 560


39. Samyal, A.; Bakhtiyar, Y.; Verma, A.; Langer, S., Studies on the seasonal variation in lipid composition of muscles, hepatopancreas and ovary of freshwater prawn, Macrobrachium dayanum (Henderson) during reproductive cycle. Advance Journal of Food Science and Technology 2011, 3, (3), 160-164. 40. Maeder, V.; Escher, B. I.; Scheringer, M.; Hungerbühler, K., Toxic ratio as an indicator of the intrinsic toxicity in the assessment of persistent, bioaccumulative, and toxic chemicals. Environmental Science & Technology 2004, 38, (13), 3659-3666. 41. Syed, N. I.; Harrison, D.; Winlow, W., Respiratory behavior in the pond snail Lymnaea stagnalis. J Comp Physiol A 1991, 169, (5), 541-555. 42. Sutcliffe, D. W., Quantitative aspects of oxygen uptake by Gammarus (Crustacea, Amphipoda): a critical review. Freshwater Biology 1984, 14, (5), 443-489. 43. Lukančič, S.; Žibrat, U.; Mezek, T.; Jerebic, A.; Simčič, T.; Brancelj, A., Effects of exposing two non-target Crustacean species, Asellus aquaticus L., and Gammarus fossarum Koch., to atrazine and imidacloprid. Bull Environ Contam Toxicol 2010, 84, (1), 85-90. 44. Tsunoyama, K.; Gojobori, T., Evolution of nicotinic acetylcholine receptor subunits. Molecular Biology and Evolution 1998, 15, (5), 518-527. 45. Snyder, M. J., Aquatic P450 Species. In The Ubiquitous Roles of Cytochrome P450 Proteins, John Wiley & Sons, Ltd: 2007; pp 97-126. 46. Lind, R. J.; Clough, M. S.; Reynolds, S. E.; Earley, F. G. P., [3H]Imidacloprid labels high- and low-affinity nicotinic acetylcholine receptor-like binding sites in the aphid Myzus persicae (Hemiptera: Aphididae). Pesticide Biochemistry and Physiology 1998, 62, (1), 3-14.

561



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562

Table 1 Toxicokinetic parameters and total body bioaccumulation factors (BAFs) ± standard errors

563

for each combination of species and chemical. RPF denotes retention potential factor for diazoxon,

564

the toxic biotransformation product of diazinon and MEF is the metabolite enrichment factor. Parameter Diazinon kin kout diazinon kout diazoxon kact BAF (parent) RPF (diazoxon) MEF (diazoxon) Imidacloprid kin kout BAF Propiconazole kin kout BAF

Unit

Lymnaea stagnalis

Gammarus pulex

Gammarus fossarum

L·kg-1·d-1 1/d 1/d 1/d L/kg L/kg

116.7 ± 5.252 0.204 ± 0.040 6.937 ± 13.28 0.114 ± 0.009 367.5 ± 79.54 0.016 ± 0.031 6.020 ± 11.61

118.9 ± 16.5634 8.464 ± 1.50634 3.278 ± 0.50734 0.896 ± 0.11334 12.70 ± 3.287 0.273 ± 0.055 3.472 ± 1.135

99.98 ± 6.848 0.000 ± 0.221 121.6 ± 72.29 6.007 ± 0.425 16.64 ± 1.639 0.049 ± 0.030 0.822 ± 0.499

L·kg-1·d-1 1/d L/kg

12.55 ± 1.255 2.796 ± 0.308 16.86 ± 2.511

1.960 ± 0.06733 0.267 ± 0.02733 7.341 ± 0.784

4.066 ± 0.239 0.241 ± 0.032 7.341 ± 1.079

L·kg-1·d-1 1/d L/kg

171.3 ± 13.52 0.458 ± 0.067 374.1 ± 62.45

99.90 ± 15.65 4.303 ± 0.655 23.21 ± 5.071

103.5 ± 9.552 4.559 ± 0.444 22.71 ± 3.046

-

565


Page 25 of 29

566


Figure legends

567 568

Figure 1 Examples of visual images (picture on top) and radioluminographs (picture below on

569

white background) of Lymnaea stagnalis, Gammarus pulex and Gammarus fossarum exposed to

570

diazinon, imidacloprid and propiconazole. Red indicates the highest concentration in the

571

radioluminographs and blue the lowest (yellow and green as gradient between these two extremes).

572

Figure 2 Imidacloprid, propiconazole, diazinon and diazoxon distribution in the tissues of Lymnaea

573

stagnalis, Gammarus pulex and Gammarus fossarum given as bioaccumulation factors (BAF and

574

MEF for diazoxon). Total body BAFs and MEFdiazoxon were obtained from toxicokinetic model

575

parameters quantifying the bioaccumulation in steady state while the tissue specific values were

576

calculated as concentration in tissues/concentration in media. The tissue MEFdiazoxon were calculated

577

from total diazinon values obtained from the QWBA analysis by multiplying these values with the

578

ratio of the parent compound or metabolite to the total body burden, resulting in a ‘24h pseudo

579

MEF’ (not at steady-state). See text for more details on calculating the BAFs in sections Methods,

580

Bioaccumulation and Chemical distribution.

581

Figure 3 Lethal exposure concentrations (LC50) after four days of exposure, internal lethal

582

concentrations based on total body burden (ILC50) and internal lethal concentrations based on

583

concentrations in the target tissues (ILC50 target) in Lymnaea stagnalis, Gammarus pulex and

584

Gammarus fossarum. As propiconazole is expected to act via narcosis, it affects cell membranes in

585

all tissues. Therefore, the ILC50 target was based on bioaccumulation factors in muscle tissue

586

representing internal tissues in general. The ILC50 target values for the neurotoxicants were created

587

using BAFs (imidacloprid) or MEFs (diazinon/diazoxon) for the nervous system. Mean values for

588

known baseline toxicants (solid grey line) and their range (dashed grey lines)13, 40 are indicated in

589

the imidacloprid and propiconazole plots. The standard error (SE = Standard deviation / Mean) was 25 ACS Paragon Plus Environment


Page 26 of 29

590

calculated to quantitatively indicate the differences in toxicity values among the species.

591

Imidacloprid results for G. fossarum need to be compared with caution because the values given are

592

based on a LC50 for 6 days, not 4 days like for all others, due to non-concentration dependent

593

mortality until day 6 (see SI).

594 595


Page 27 of 29


596

597 598

Figure 1

599



Page 28 of 29

600 601

Figure 2

602


Page 29 of 29


603 604

Figure 3


Importance of Toxicokinetics for Interspecies Variation in Sensitivity to

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