Biological responses to phenylurea herbicides in fish and amphibians: New directions for characterizing mechanisms of toxicity

    Biological responses to phenylurea herbicides in fish and amphibians: New directions for characterizing mechanisms of toxicity Vicki L. Marlatt, Christopher J. Martyniuk PII: DOI: Reference:

S1532-0456(17)30002-9 doi:10.1016/j.cbpc.2017.01.002 CBC 8274

To appear in:

Comparative Biochemistry and Physiology Part C

Received date: Revised date: Accepted date:

18 August 2016 11 January 2017 13 January 2017

Please cite this article as: Marlatt, Vicki L., Martyniuk, Christopher J., Biological responses to phenylurea herbicides in fish and amphibians: New directions for characterizing mechanisms of toxicity, Comparative Biochemistry and Physiology Part C (2017), doi:10.1016/j.cbpc.2017.01.002

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ACCEPTED MANUSCRIPT Biological responses to phenylurea herbicides in fish and amphibians: new directions for characterizing mechanisms of toxicity

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Vicki L. Marlatta*# and Christopher J. Martyniukb,c# a

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Department of Biological Sciences, Simon Fraser University, 8888 University Drive, Burnaby, British Columbia, Canada b

c

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Department of Physiological Sciences and Center for Environmental and Human Toxicology, UF Genetics Institute, College of Veterinary Medicine, University of Florida, Gainesville, Florida 326111 USA Canadian Rivers Institute, Canada.

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#These authors contributed equally to the study

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*Corresponding author: Vicki L. Marlatt, Department of Biological Sciences, Simon Fraser University, 8888 University Drive, Burnaby, BC, Canada, V5A 1S6, Ph: 778-782-4107, Email: [email protected]

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ACCEPTED MANUSCRIPT Abstract Urea-based herbicides are applied in agriculture to control broadleaf and grassy weeds, acting to either inhibit photosynthesis at photosystem II (phenylureas) or to inhibit acetolactate

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synthase acetohydroxyacid synthase (sulfonylureas). While there are different chemical

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formulas for urea-based herbicides, the phenylureas are a widely used class in North America and have been detected in aquatic environments due to agricultural run-off. Here, we summarize

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the current state of the literature, synthesizing data on phenylureas and their biological effects in

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two non-target animals, fish and amphibians, with a primary focus on diuron and linuron. In fish, although the acutely lethal effects of diuron in early life stages appear to be >1 mg/L, recent

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studies measuring sub-lethal behavioral and developmental endpoints suggest that diuron causes adverse effects at lower concentrations (i.e. <0.1 mg/L). Considerably less toxicity data exist for

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amphibians, and this is a knowledge gap in the literature. In terms of sub-lethal effects and mode of action (MOA), linuron is well documented to have anti-androgenic effects in vertebrates,

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including fish. However, there are other MOAs that are not adequately assessed in toxicology studies. In order to identify additional potential MOAs, we conducted in silico analyses for linuron and diuron that were based upon transcriptome studies and chemical structure-function relationships (i.e. ToxCast™, Prediction of Activity Spectra of Substances). Based upon these analyses, we suggest that steroid biosynthesis, cholesterol metabolism and pregnane X receptor activation are common targets, and offer some new endpoints for future investigations of phenylurea herbicides in non-target animals.

Keywords: embryos, fish, amphibian, agriculture, comparative transcriptomics, chemical structure-function relationships 1. Current Use and Presence of Phenylurea Herbicides in North America

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ACCEPTED MANUSCRIPT The widespread use of pesticides globally has resulted in their frequent detection as contaminants in surface and drinking water, and has heightened concern regarding the

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bioaccumulation, acute and chronic health risks of pesticide exposure in humans and wildlife. In

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Canada for example, more than 500 pesticides are registered and, like most countries, there is no central registry of pesticide sales or use data. However, government and peer-reviewed scientific

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studies frequently report the presence of pesticides as individual compounds or mixtures in the

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environment (Brimble et al., 2005). In a recent national survey of 141 pesticides and transformation products in surface and ground waters in Canada, 102 were detected across the

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country (Environment Canada, 2011). One class of widely used herbicides applied via ground and aerial equipment in North America for control of broadleaf and grass weeds during pre- and

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post-emergent crop production is the urea-based herbicides, with linuron and diuron the most commonly used and studied in this class.

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Diuron is registered in Canada as a herbicide for grape and asparagus crops as well as non-crop areas (ponds, irrigation and drainage ditches) and is also present in non-pesticide products industrial in nature (e.g. hardening agents in epoxy resins, curing agent in epoxy adhesives for bonding of metal agents and manufactured industrial items used in the transportation industry) (Health Canada and Environment Canada, 2011). In the US, diuron was first registered in 1967 by the United States Environmental Protection Agency (USEPA; under the Federal Insecticide, Fungicide and Rodenticide Act) for use on a wider range of agricultural crops and similar non-agricultural use sites compared to Canada, and is often used in combination with other herbicides and surfactants (USEPA, 2003). Specifically, in the US, diuron is used as a pre-emergent herbicide to control weeds in crops such as alfalfa, artichoke, asparagus, bananas, barley, Bermuda grass pastures, blueberries, cranberries, gooseberries, corn, 3

ACCEPTED MANUSCRIPT cotton, grapes, perennial grass-seed crops, papayas, peppermint, pineapple, plantains, sorghum, sugarcane, small grains, and several fruit- and nut-tree crops as well as certain ornamentals

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(USEPA, 2003). In the US between 2001 and 2007, the use of diuron was estimated at 2 to 6

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million pounds annually and was ranked as one of the most commonly used conventional pesticide active ingredients (Grube et al., 2011). Linuron on the other hand is less restricted in

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use compared to diuron, but has recently been re-evaluated by Health Canada and is slated to be

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phased out due to human health and environmental health concerns (Health Canada, 2012). In Canada, linuron use has been registered by Health Canada under the Pest Control Products Act

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to control a variety of weeds on field and garden crops such as corn, carrots, potatoes, fruit trees, wheat oats and barley since the early 1990s (Health Canada, 2012). In the US, linuron was first

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registered in 1966 for a wide range of agricultural crops (carrots, celery, asparagus, soybean, corn, potatoes) and non-food crop agricultural uses (ornamental bulbs, poplar trees for use in

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shelterbelts in the mid-west) and a registration review by the USEPA is currently underway (USEPA, 1995, 2010). Although plants are the intended target and most sensitive to the acute toxic effects of phenylurea herbicides, there has been a steady gathering of information (e.g. toxicity, fate, distribution, chemistry) about phenylurea herbicides over the past twenty years (average 60-70 studies/year) in a variety of taxa (Supplemental Figure 1). Indeed, the relatively low acute toxicity of phenylurea herbicides in mammals, birds and fish compared to plants, algae and invertebrates is evident, but studies demonstrating chronic toxicity of these compounds in vertebrates are mounting (Hogan et al., 2012; Marlatt et al., 2013; Martyniuk et al., 2012; Pereira et al., 2015; Turner et al., 2003; Uren Webster et al., 2015; USEPA, 2015). However, our knowledge regarding their non-target effects in some animals, particularly fish and amphibians is very limited. In fact, studies that investigate the effects of these herbicides in fish and

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ACCEPTED MANUSCRIPT amphibians only comprise a small fraction of the total number of publications for these compounds (Supplemental Figure 2). Therefore, the objective here was to review the available

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scientific literature focused on fish and amphibians and conduct vertebrate-based in silico

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analyses of phenylurea-based herbicides to determine the scope of non-target adverse effects in

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fish and amphibians, and potential conserved MOAs of these environmental contaminants. 2. Chemical structures

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Urea-based herbicides are a broad class of herbicide that contain urea (CO(NH2)2) within

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an organic compound (Lanyi & Dinya, 2005). Urea-based herbicides include classes such as sulphonylureas (general structure of R1–NH–C(O)–NH–SO2–) (e.g. nicosulfuron, rimsulfuron),

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benzoylureas (BU) (general structure of R1-NH–C(O)–NH–-CO-substituted phenyl) (e.g.

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diflubenzuron, luferuron), thiadiazolylurea herbicides (e.g. buthiuron, thiazafluron), triazinylsulfonylurea herbicides (e.g. chlorsulfuron, iofensulfuron , triasulfuron, tribenuron) and

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phenylureas (Niessen et al., 2010). Phenylurea herbicides have a general structure of phenyl– NH–C(O)–NR2, thus there are two nitrogen groups joined by a carbonyl (C=O) functional group, with one of the N-groups containing a phenyl group (C6H5) and the other consisting of various substituted moieties (Morais et al., 2011). There are more than 20 phenylurea based chemicals in use for plant control (e.g. anisuron, chloroxuron, diuron, linuron, isoproturon, monuron, neburon, phenobenzuron, siduron, tetrafluron among others), and structures compiled from the EMBL-Chemical Entities of Biological Interest (ChEBI) reveal that there are a number of chemical substitutions that include fluorine, chlorine, and methyl groups as well as other moieties (Figure 1). As stated above, two of the most widely used and studied phenylurea herbicides include diuron and linuron (located in the boxes in Figure 1). Diuron (CAS Type 1 Name: Urea, N'-(3,45

ACCEPTED MANUSCRIPT dichlorophenyl)-N,N-dimethyl-; CAS Registry Number 330-54-1) and linuron (CAS Type 1 Name Urea, N'-(3,4-dichlorophenyl)-N-methoxy-N-methyl-; CAS Registry Number 330-55-2)

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both have two chlorine groups associated with the phenyl ring, but at different positions.

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Another difference between the two is that diuron contains a methyl group on a substituted nitrogen whereas linuron contains a methoxyl group. This review focuses on diuron and linuron

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due to their relatively abundant use in agriculture compared to other urea-based herbicides and

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summarizes what is known about their fate, distribution, and biological effects in fish and

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amphibians. 3. Environmental fate

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Table 1 summarizes some of the key physical/chemical and persistence data relevant to

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the environmental fate of diuron and linuron. Aerobic and anaerobic microbial metabolism are

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the main degradation pathways for both diuron and linuron in soil and water (Table 1; USEPA, 2003; USEPA, 1995; Jewett & Koper, 2016). Degradation appears to be more rapid in aquatic environments than terrestrial environments (Table 1). Indeed, for diuron lab-based soil microbial metabolism studies have demonstrated half-lives of 372 and 1000 days under aerobic and anaerobic conditions, respectively, in silt loam soils compared to <34 days in aerobic and anaerobic aquatic microbial metabolism studies (Hausmann, 1992; Hausmann and Kraut, 1992; Hawkins et al., 1990; USEPA, 2003; Yu, 1988). If released into the soil, diuron is expected to have low to moderate adsorptivity to soil based on its organic carbon-water partition coefficient (experimental Koc = 2.4±0.2; (Thomas et al., 2002)). If released into water, diuron is expected to mainly reside in water (99%; water solubility 42 mg/L at 25 °C) and to a very small extent partition into sediment (<1%) (DuPont, 1989). Diuron has a low volatility based on its low vapour pressure (1.1 x 10-7 Pa at 25 °C) and low Henry’s law constant (7.04 x 10-6 Pa·m3/mol; 6

ACCEPTED MANUSCRIPT (Tomlin, 2005-2006). Consequently, volatilization is predicted to be insignificant in water and soil, except under several days or weeks of hot, dry conditions particularly when it resides on the

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surface of soil (Health Canada and Enviornment Canada, 2011; USEPA, 2003). Several major

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(>10% of the applied parent) and minor (<10% of the applied parent) metabolites have been reported in soil microbial degradation studies. For example, in well-oxygenated brown

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calcareous soil under field conditions after 50 days over 80% of the parent compound diuron was

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still present, but the following major metabolites were also produced: (1) N-(3,4dichlorophenyl)-N-methylurea; 3,4-dichlorophenylurea; and (3) dichloroaniline (Howard, 1991).

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Linuron exhibits characteristics similar to those observed for diuron (i.e. low volatility, similar persistence in soil and water), however, linuron is ~twice as soluble in water (81 mg/L at 25 °C)

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(Jewett and Koper, 2016; Kidd and James, 1991; Monson, 1986). Biodegradation of studies of linuron investigating microbial metabolism in aerobic soil report a half-life ranging from 52 to

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1100 days depending on the soil type (Schneiders, 1990; USEPA, 1995), which is within the range reported for diuron. The half-life of linuron in aerobic water systems due to microbial metabolism is very similar to diuron as well (i.e. than 42 days), and several major (desmethoxy linuron, norlinuron and dichloroaniline) and minor metabolites (desmethyl linuron, norlinuron and other unknowns) of linuron have been reported (Jewett and Koper, 2016). There is a paucity of environmental fate and biological effects data regarding the metabolites of diuron and linuron produced via biodegradation pathways in soil and water. Although limited experimental data for bioaccumulation and bioconcentration of phenylurea herbicides exists, the few studies conducted to date for diuron and linuron suggest low potential for these phenomena in fish. In particular, the bioconcentration factors (BCFs) and bioaccumulation factors (BAFs) reported indicate values less than 100 for diuron and less than 7

ACCEPTED MANUSCRIPT 2400 for linuron, (Call et al., 1987a; CHRIP, 2008; Francis et al., 1985; Kenaga, 1980; Pont, 1984; Tucker et al., 2003). For example, bioconcentration factors (BCFs) in bluegill sunfish

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during 28 day exposure experiments to waterborne 0.1 and 1.0 mg/L radiolabelled linuron

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resulted in BCFs for whole fish of 49, 240 for viscera, 34 for muscle and 39 for carcass tissues (Du Pont Canada, 1984) (Du Pont 1984). Metabolites present after 28 days were identified as

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desmethyl linuron, norlinuron, and glucuronide conjugates (Du Pont, 1984). A toxicity, uptake

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and elimination study in fathead minnows showed that diuron did not accumulate in fish tissue to a large extent with nearly 99% elimination after 21 days and a bioconcentration factor of 2 (Call

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et al., 1987a). Four metabolites were identified during this study, one identified as 3,4dichloroaniline, two additional demethylated compounds that were not positively identified and

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the fourth metabolite was not identified (Call et al., 1987a). These low BCFs and BAFs in fish coincide with the low octanol-water partition coefficients (log Kow ~2.5-3.0) and moderate-high

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water solubility for both of these phenylureas (Table 1). There is an apparent absence of studies reporting BCFs or BAFs in amphibians in the currently available scientific literature. 4. Presence in Aquatic Environment Contamination of the environment by pesticides occurs as a result of direct applications or indirectly via spray drift, leaching, runoff events and dry/wet deposition events (CCME 1999). Although there is a paucity of data worldwide reporting surface and ground water concentrations of phenylurea herbicides, linuron and diuron have been reported in the ng/L to low µg/L range in a number of studies (Berryman and Giroux, 1994; Environment_Canada, 2011; Frank et al., 1987; Gatidou et al., 2007; Kotrikla et al., 2006; O'Neill and Bailey, 1987; Schuler and Rand, 2008; USEPA, 1992; USGS, 1998; Woudneh et al., 2009; Xing et al., 2012). A large study in the US conducted by the US Geological Survey (USGS) National Water Quality Assessment 8

ACCEPTED MANUSCRIPT Program (NAWQA) collected 1420 surface water samples from 62 agricultural stream sites during the period from 1992-1998, and diuron was detected in 104 (7.3%) of the samples

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(detection limit = 0.05 µg/L) with a maximum concentration of 13 µg/L (USGS, 1998). Similar

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concentrations of diuron have also been reported in ground water in the US (USEPA, 1992; USGS, 1998). For example, the range of ground water concentrations measured for diuron was

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0.34 to 5.37 µg/L in a USEPA survey of ground water in California, Florida, Georgia and Texas

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(USEPA, 1992). Reports of environmental concentrations of linuron in the US are similar to those reported for diuron. A recent study reported concentrations of linuron in a Florida stream

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receiving agricultural run-off of 4.42 µg/L (Schuler and Rand, 2008). In USGS nation-wide surveys in the late 1990s and early 2000’s, linuron was detected in 2.7% of 5196 surface water

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samples collected from 40 agricultural impacted streams (detection limit of 0.01 µg/L) and a maximum concentration of 1.4 µg/L was reported (Davy and Shaugnessy, 2008). Linuron has

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also been detected in ground water in four states including Georgia, Missouri, Virginia, and Wisconsin at levels ranging up to 5 µg/L (Hoheisel et al., 1992). In Canada, environmental levels of linuron have been reported more frequently than diuron, but the concentrations observed for both of these herbicides appear to be similar to the observations in the US. The first National Water Quality Surveillance Program (NWQSP) under Environment Canada was conducted during 2003-2005, and reported linuron as one of the top five pesticide active ingredients sold in the Atlantic provinces of Prince Edward Island (PEI), Nova Scotia (NS) and New Brunswick (NB) (Environment Canada, 2011). Interestingly, during this survey, linuron was only detected in 16% of surface water samples in NB and was not detected in any of the 82 and 48 samples collected in 2003-2005 in PEI and NS, respectively. This was in spite of relatively sensitive measurement method detection limits (0.00006-0.003 9

ACCEPTED MANUSCRIPT µg/L) and collections in areas impacted by agriculture. However, a subsequent study in 2006 and 2007 showed an increase in detection frequency in the Atlantic provinces with the detection

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of linuron in 43.6% of samples (mean concentration 1.83 µg/L) and 39.8% of samples (mean

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concentration of 1.81 µg/L), respectively (Xing et al., 2012). These reports may suggest that pesticide sales data are of limited value when trying to capture information on recent/current

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pesticide use and their residues in the environment. In western Canada in British Columbia (BC)

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during the NWQSP 2003-2005 survey, linuron was also detected relatively frequently compared to other herbicides. In particular, 33 of 93 (35.5%) surface water samples collected in

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agricultural areas in BC (Lower Fraser Valley and Okanagan Basin) between 2003-2005 contained linuron at concentrations ranging from 0.00041-1.05 µg/L (Health Canada and

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Environment Canada, 2011). In addition, linuron was also detected in isolated lakes not impacted by agriculture in Ontario in 19 of 163 (11.7%) samples at low levels (ranged from

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<0.000046 – 0.00410 µg/L; Environment Canada, 2011). Limited data for diuron in Canada is available, and during the NWQSP survey, it was only measured in one province in Canada, Quebec (QB). During this 3 year national survey (2003-2005) diuron was not detected in 2 of the 3 years in a total of 121 surface waters collected in agriculturally impacted water bodies in QB, however the method detection limit was relatively high (varied between 0.080-0.250 µg/L; Environment Canada, 2011). Of 62 surface water samples in the third year of this survey, only one sample was reported to contain 0.270 µg/L of diuron (method detection limit of 0.250 µg/L; Environment Canada, 2011). Published data of linuron detection in ground water in ng/L to low µg/L concentrations in several provinces in Canada is available (Health Canada and Enviornment Canada, 2011), but no reports describing diuron measurements in Canadian ground water were identified at the time of this review.

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ACCEPTED MANUSCRIPT The presence of linuron and diuron in surface and ground waters in North America in several studies concurs with the water-soluble nature of these compounds. However, the

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detection of linuron in isolated lakes combined with the detection of linuron in precipitation in

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the Canadian NWQSP survey in ON and BC ranging from <0.00009 to 0.147 µg/L suggest volatilization of linuron and long-range transport from areas of pesticide use to pristine areas is

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occurring (Health Canada and Environment Canada, 2011). These phenomena were not

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predicted based on the physical or chemical characteristics of linuron, and are currently poorly understood and future environmental monitoring and lines of investigation should take this into

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consideration. Indeed, many of the environmental monitoring studies were conducted in the 1990’s and early 2000’s and the frequency of sampling and length of sampling periods are

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limited. Ongoing environmental monitoring to produce larger and more comprehensive data sets to better understand the spatial and temporal occurrences of the phenylurea herbicides, and

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pesticides in general, will be of great benefit to understanding their environmental fate and potentials risks to humans and the environment. 5. Mechanism of action in plants, frogs, and fish Photosynthesis is the complex process of converting sunlight into usable energy by a plant and normal function of the multiple enzymes involved in this conversion is essential for life, and thus this process is an ideal target for herbicides. The chloroplasts are the organelle that harnesses the energy from the sun, and electrons are passed along photosystems using donor molecules where they are ultimately used to generate chemical energy in the form of adenosine triphosphate (ATP). Many herbicides have been designed to act on different points in the photosynthetic pathway, for example, impeding energy production as energy transfer inhibitors, electron transport inhibitors, or uncouplers of ATP production. In plants, phenylurea herbicides 11

ACCEPTED MANUSCRIPT act primarily to inhibit photosynthesis at Photosystem II, and are considered to be electron transfer inhibitors (Moreland, 1980). Photosystem II is the first protein complex in light-

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dependent reactions in the chloroplast, thus these compounds block ATP production early on in

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the process and inhibit growth. Diuron for example has been shown to interact with proteins such as the QB protein in chloroplasts, resulting in allosteric inhibition (Van Rensen, 1989). To

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complicate matters, multiple ureic-based herbicides can affect Photosystem II and other proteins

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in the electron transport chain in an additive or synergistic manner, compounding the overall impacts on growth and survival of the plant (Redondo-Gomez et al., 2007). It is important to

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point out that there is diversity in the primary mode of action of many urea-based herbicides and there are other primary targets within the chloroplast that can be affected by urea-based

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herbicides (Pang et al., 2003; Stidham, 1991). For example, chemicals within the sulfonylurea class can inhibit acetolactate synthase (or acetohydroxyacid synthase), an enzyme that catalyzes

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the first step in the synthesis of the branched-chain amino acids (Pang et al., 2003). 5.1 Biological effects on non-target aquatic species The rationale for targeting light dependent ATP production in the plant chloroplast using herbicides is clear, as this process is not found in species in the Kingdom Fungi and Animalia. However, this does not preclude non-target effects from these herbicides in species that do not contain chloroplasts. Indeed there appears to be a wide range of biological effects that have been reported in both aquatic and terrestrial organisms, the most prevalent being that of androgen receptor antagonism in fish and mammals (Hogan et al., 2012; Hogan et al., 2008; Jolly et al., 2009; USEPA, 2015; Wolf et al., 1999). It is also expected that there can be adverse biological effects of these compounds that are not yet identified. Moreover, reports of several metabolites of phenylurea herbicides such as linuron and diuron in microbial, fish and mammalian laboratory

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ACCEPTED MANUSCRIPT toxicity tests exist, however, data on the toxicity of these metabolites and their prevalence in the environment are scarce. As both mitochondria and chloroplasts function to shuttle electrons

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across electrochemical gradients using water and oxygen, it is reasonable to suggest that these

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herbicides may also affect mitochondrial bioenergetics and the generation of ATP in non-target species. To synthesize the literature on the adverse effects of phenylurea herbicides in fish and

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amphibians, we performed a comprehensive literature search on the toxicity of a number of

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phenylurea parent compounds. The majority of the studies obtained in the search were those

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investigating diuron and linuron and these are summarized below. 5.1.1 Literature search method and results

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To determine the scope of adverse effects of phenylurea-based herbicides on fish and

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amphibians, a literature search was undertaken using Web of Science and Google Scholar.

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Search terms were [product] and [taxon], where [product] was either anisuron, buturon, chlorbromuron, chloreturon, chlorotoluron, chloroxuron, daimuron, difenoxuron, dimefuron, diuron, fenuron, fluometuron, fluothiuron, isoproturon, linuron, methiuron, methyldymron, metobenzuron, metobromuron, metoxuron, monolinuron, monuron, neburon, parafluron, phenobenzuron, siduron, tetrafluron or thidiazuron; and [taxon] was either fish*, frog* and amphibian*. In addition, specific searches were made on a few common toxicity test species (e.g. rainbow trout) and by following up references cited in the publications found by the search. The review also draws heavily on recently published reports by the USEPA, USGS, Health Canada and Environment Canada for the environmental fate and presence information for diuron and linuron. Several industry studies that were not formally published but were part of product approval processes were reviewed by these government regulatory bodies and, therefore, were included in the environmental fate and presences section. However, the primary and peer13

ACCEPTED MANUSCRIPT reviewed literature on the toxicity of phenylureas on fish and frogs was the emphasis here. The following information was the acceptance criteria for each of the open peer-reviewed

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literature studies on the toxicity of phenylurea herbicides in fish and frogs prior to their inclusion

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in this review: appropriate controls; replicate treatments; ≥2 concentrations (nominal/actual) of

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active ingredient tested and stated; single toxicant exposure; and, duration of the exposure clearly provided. The results of the primary peer-reviewed literature search using Web of Science and

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Google Scholar web searches for fish and [product] were as follows: 1 acceptable and 6

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unacceptable studies for linuron; 6 acceptable and 16 unacceptable studies for diuron; and, 1 unacceptable study for fenuron, monuron and neburon. For the frog and [product] web search

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the results were as follows: 2 acceptable for diuron; 1 acceptable for isoproturon; and, 1

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acceptable for linuron.

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5.1.2 Effects of Linuron and Diuron in Fish Acute toxicity data for phenylurea herbicides in adult teleost fish is available for linuron and diuron for several species. These short-term tests are conducted during a small part of the lifecycle of an organism and the test endpoint is often death, typically expressed as an LC50 or EC50 (concentration of a chemical that caused mortality/adverse effects in 50% of the test population after a specific exposure time). In contrast, few studies examining chronic exposure scenarios and the early life stages of teleost fishes have been conducted. Here, we present a general overview of toxicity data used for guideline derivation in Canada and the US and some of the more recent developments including the testing of linuron in endocrine screening bioassays and early life stage fish studies for linuron and diuron. For adult teleosts, technical grade linuron 96-h LC50 values range from 3.2 – 16.4 mg/L 14

ACCEPTED MANUSCRIPT for rainbow trout, and values of 2.9, 5.2, and 16.2 have been obtained for channel catfish (Ictalarus punctatus), brown bullheads (Ictalarus nebulosus), and bluegill sunfish (Lepomis

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macrochirus), respectively (Du Pont Canada, 1973, 1986; Linders et al., 1990; Lysak and

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Marcinek, 1972). The Canadian Council of Ministers of the Environment (1999) derived the current Canadian Water Quality Guideline for linuron of 7.0 µg/L for the protection of aquatic

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life based on chronic toxicity in plants (CCME, 1999). Specifically, growth inhibition

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experiments in duckweed demonstrated a lowest acceptable effect concentration, a 5-d EC50 value of 70 µg/L (Stephenson, 1984), and this value was multiplied by a safety factor of 0.1

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(CCME, 1999). In contrast, no National Recommended Water Quality Criteria for Aquatic Life exists for linuron in the US and it is currently undergoing review for reregistration by the

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USEPA. However, the USEPA does state that linuron is considered slightly to moderately toxic with LC50s for acute toxicity in aquatic organisms falling within the range of 1-100 mg/L

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(USEPA, 2003). In addition, linuron was recently tested in the USEPA’s Endocrine Disruptor Screening Programs Tier 1 assay battery for adverse effects on estrogen, androgen and thyroid pathways which includes a weight of evidence analysis (USEPA, 2015). The USEPA concluded that there is consistent complementarity and redundancy of evidence based on several bioassays that linuron acts as an anti-androgen both in vitro and in vivo, and additional testing in fish models in the longer term Tier 2 assay battery is recommended. For example, in vitro linuron binds competitively to the androgen receptor (AR) and activates transcriptional activity in multiple fish species (i.e. fathead minnow, rainbow trout) and in mammals (i.e. rats and humans; reviewed in USEPA, 2015). In vivo evidence supporting anti-androgenic effects of linuron are based on several studies in rats showing adverse effects on androgen-mediated tissues or processes (i.e. Hershberger assay, pubertal assays and reproductive developmental assays)

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ACCEPTED MANUSCRIPT combined with inhibition of androgen-induced spiggin production by female stickleback fish kidney cells after linuron exposures at 150 µg/L (Hogan et al., 2012; Hogan et al., 2008; Jolly et

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al., 2009). Although some positive results have been observed in estrogen-receptor mediated

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responses in fish and other taxa (i.e. mammals), the conclusion of the USEPA weight of evidence approach was that the evidence was not sufficient to characterize linuron as estrogenic/anti-

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estrogenic. Furthermore, although the USEPA also concluded that there is sufficient evidence of

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potential interaction of linuron with the thyroid pathway in mammalian models, studies of linuron perturbing the thyroid pathway in fish and amphibians is currently completely absent

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from the literature. To date, whether environmentally relevant concentrations of linuron are adversely impacting androgen and thyroid-mediated processes in fish and frogs is unclear, and

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further studies identifying the risks of low-level exposure over multiple generations on sub-lethal adverse effects are warranted. Furthermore, whether other phenylurea herbicides, such as

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diuron, exhibit similar endocrine disrupting effects has not been tested. Several acute toxicity studies in fish have been conducted for diuron, although few chronic studies in juveniles/adults have been reported and currently no Canadian or US water quality guidelines exist for this herbicide. The acute toxicity of diuron appears to be similar to linuron in fish. In juvenile/adult freshwater teleosts, data collected from acute toxicity studies report LC50 values ranging from 0.71 – 14.2 mg/L, with cutthroat trout the most sensitive teleost species followed by lake trout (Salvelinus namaycush), rainbow trout, coho salmon (Oncorhynchus kisutch), bluegill sunfish (Lepomis macrochirus) and the least sensitive fathead minnow tested (Call et al., 1987b; OPP Pesticide Ecotoxicity Database, 2008; Mayer and Ellersieck, 1986; Okamura et al., 2002; USEPA, 2003). Few studies examining the effects of phenylurea herbicides on the early life stages of 16

ACCEPTED MANUSCRIPT teleosts have been conducted, and of these most have tested diuron (Table 2). Although the lethal effects of diuron in embryos, larvae and juveniles appears to fall within the >1-100 mg/L

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range similar to the potency observed in adult teleost studies to date, several studies measuring

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sub-lethal behavioural and developmental endpoints show adverse at concentrations <0.1 mg/L (Table 2). For example, behavioural experiments in juvenile goldfish after 1 day exposures to

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0.0005, 0.005 and 0.05 mg/L diuron demonstrated significant effects on burst swimming at 0.05

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mg/L, and significantly decreased grouping behaviour at 0.005 mg/L although not at the highest concentration tested (Saglio and Trijasse, 1998). A study in pink snapper (Pagnus auratus)

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suggests short-term exposure to 0.05 mg/L can have detrimental effects on hatching success and increased deformity rates in larvae during acute exposures initiated during embryogenesis and

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continued to shortly after hatching (Gagnon and Rawson, 2009). These effects on hatching success suggest that diuron is capable of crossing the egg barrier and impacting embryogenesis.

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Although no acute toxicity studies in the fathead minnow model reporting developmental effects were obtained, similar potency of diuron on development endpoints was observed in early life stage fathead minnow 60 day exposures whereby significant increased gross deformities and decreased survival were observed at 0.078 mg/L (Call et al., 1987a). Interestingly, studies in embryonic Golden medaka (Oryzias latipes) prior to hatching demonstrate that this species is considerably less sensitive than pink snapper and the fathead minnow early life stages (Newman et al., 2001), but does further support the notion that diuron crosses the egg barrier and adversely impacts embryonic development. In particular, Newman et al. (2001) conducted sub-lethal waterborne 15 day exposures in Golden medaka that were initiated in embryos (early to late blastula) and continued through to post-hatched larval developmental stages, and observed hatching delays at 2.01 mg/L and 3.96 mg/L and various morphological abnormalities (i.e. yolk

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ACCEPTED MANUSCRIPT sac edema, failed swim bladder inflation and gall bladder enlargement) at 7.50 and 15.15 mg/L. Although the studies on early life stages of fish to date are few in number and are mainly focused

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on diuron, they do represent a wide range of experimental designs and test species. And, based

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on the majority of the environmental concentrations of phenylurea herbicides measured to date (i.e. typically < 0.015 mg/L), it appears as though biologically significant sub-lethal effects on

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development and behaviour reported to date in fish occur at ~5-10 fold higher concentrations

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than what is presently found in surface waters.

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5.1.3. Effects of Diuron and Linuron in Amphibians

In contrast to fish, there is a paucity of research on amphibians and the relative toxicity

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of phenylurea herbicides. Studies that do report on these phenylurea herbicides and their effects in amphibians are primarily confined to the field and occur in agricultural regions with high

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pesticide usage. As such, there are often confounding variables that are related to adverse effects observed in local amphibian populations. For example, hatching success of Great Basin spadefoot eggs was negatively correlated with increasing pesticide and nutrient concentrations in ponds in the South Okanagan Valley, British Columbia (CAN), and it was determined that eggs exposed to a mixture of pesticides affected amphibian development by additive or synergistic effects with other environmental factors (Bishop et al., 2010). Linuron ranging 70-120 ng/L was detected in these ponds at both the reference sites and sprayed orchards, and there appeared to be no relationship between this herbicide and hatching success across spatial and temporal sampling events.

In an effort to address the lack of data regarding the toxicity of phenylurea herbicides in amphibians, the effects of diuron in frogs living in riparian zones were investigated using acute 18

ACCEPTED MANUSCRIPT and sub-chronic (2 week) toxicity bioassays (Schuytema and Nebeker, 1998). The species investigated included the Pacific treefrog (Pseudacris regilla), bullfrog (Rana catesbeiana), red-

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legged frog (Rana aurora), and African clawed frog (Xenopus laevis) at life stages that included

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both embryos and tadpoles. Embryos were assessed for mortality following exposure to diuron, and this was only ~13% for P. regilla when exposed to 29.1 mg/L diuron. However,

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approximately 80% of the embryos showed deformities, such as edema. Lower doses tested

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(down to 0.5 mg/L) did not appear to have significant effects on the embryos when compared to the control group. The study also reported that the lowest NOAEL values for acute toxicity for P.

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regilla, R. catesbeiana, and R. aurora were comparable, ranging 7.63 - 14.5 mg/L, suggesting that these species in particular were equally sensitive to the herbicide (Schuytema and Nebeker,

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1998). Mortality in the P. regilla tadpoles at 29.1 mg/L occurred at ~62.5 - 87.5% while in R. catesbeiana, there was a wider range in % mortality that depended upon the bioassay conducted

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(10-90% at 29.1 mg/L diuron). Eighty percent of the R. aurora tadpoles died at 29.1 mg/L diuron (Schuytema and Nebeker, 1998). Interestingly, diuron had some effect on tadpole growth but it was unclear if this was an inhibitory effect or stimulatory effect in tadpoles, as values for length and wet weight among treatments diverged from controls throughout the trials. In terms of toxicity to the amphibian species, the levels of diuron that adversely affected the tadpoles were much higher than that normally used for field application or measured as contaminants in surface waters. However, the lowest concentration tested in this study was 0.5 mg/L and whether lower levels more relevant to environmental exposure scenarios would elicit sub-lethal effects in these species that are observed in other vertebrates, particularly on the endocrine system, was not incorporated into this study design.

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ACCEPTED MANUSCRIPT Greulich and colleagues (2002) studied the biological effects and uptake of isoproturon, another substituted phenylurea herbicide, to early developmental stages of both red-bellied toad

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(Bombina bombina) and the yellow-bellied toad (Bombina variegata) at environmentally

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relevant levels and did see evidence of sub-lethal effects on development, growth and behavior. Using 14C-labeled isoproturon, the researchers demonstrated that these herbicides can pass

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through the protective jelly matrix surrounding the egg leading to a direct exposure on the

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embryos (Greulich et al., 2002). Exposure to environmentally relevant concentrations of isoproturon at levels detected in agricultural ponds (~1 µg/L) resulted in up to 50% of the B.

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variegata tadpoles becoming paralyzed with high mortality (Greulich et al., 2002). Moreover, at concentrations ranging from 0.1 to 100 µg/L, there was a decrease in the ability of the tadpoles to

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dart away with a normal avoidance response when prodded. Tadpoles also showed induction of the enzyme glutathione S transferase (GST) following exposures to pure isoproturon and

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mixtures of the herbicide for applications (Greulich et al., 2002). In addition to endpoints related to toxicity, development, growth and behavior, a study in Xenopus provides evidence that endocrine disruption may also result from low-level exposures with these herbicides. In vitro incubation of cultured Xenopus laevis oocytes with diuron and linuron for 20 h resulted in elevated progesterone and decreased testosterone production, and these changes in steroids were accompanied by decreased ovulation in the ovulation assay (Orton et al., 2009). Noteworthy was that isoproturon impaired ovulation without affecting the steroid levels, suggesting that there may be multiple adverse outcome pathways underlying the decreased ovulation events. There was also evidence for anti-androgenicity for diuron (15.6-31.3 µM or 3.5-7.2 mg/L), linuron (0.98-62.5 µM or 0.25-15.6 mg/L), and isoproturon (125-250 µM or 25.7-51.5 mg/L), using a yeast screening system for receptor activation/inactivation (Orton et al., 2009). Therefore, there

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ACCEPTED MANUSCRIPT is some evidence that phenylurea herbicides as a group may act as endocrine disruptors in amphibians. This coincides with the recent USEPA conclusions for linuron as an anti-androgren,

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and screening of additional phenylurea based herbicides in the USEPA’s Endocrine Disruptor

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Screening Programs assay battery merits further attention.

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In summary, the literature available for biological effects of phenylurea herbicides on amphibians is scarce compared to mammals, fishes and invertebrates. These herbicides are

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mobile in sediment and are used in agriculture to control weeds, and run-off events are likely.

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Additional studies are needed to determine the extent to which these herbicides contribute to the worldwide decline in amphibian populations, as some reports suggest that pesticides are a

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significant contributor to this phenomenon (Bruhl et al., 2011; Bruhl et al., 2013), although

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global declines in amphibians are likely due to multiple stressors (Knapp et al., 2016).

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6. Molecular responses to linuron and diuron in fish and amphibians: What endpoints should we measure to assess exposure risk? To date, data are limited regarding the molecular signaling cascades activated or inhibited by phenylurea herbicides in aquatic organisms. In fathead minnow, ovary explants incubated with linuron at a concentration of 250 µg/L (10− 6 M) in the media for 12 hours showed changes in transcripts related to cell signaling (WNT-signaling, forkhead box A3), transcription factor activation (Pou class 2 homeobox 1, Leucine zipper), immune response (interleukin 1alpha, and interleukin 17A), and steroid hormone biosynthesis (Ornostay et al., 2013). In the liver of female fathead minnows, linuron was also observed to affect proteins related to epidermal growth factor (EGF) and MAPK14 signaling, and a molecular network that included these proteins were significantly increased following waterborne exposure to linuron (Martyniuk et al., 2012). In the same study, other examples of molecular targets of linuron included prostaglandin 21

ACCEPTED MANUSCRIPT E synthase 3 (cytosolic) (PTGES3), a protein that was increased in the liver, and ribosomal proteins (L8, L26, S24 among others) that were decreased in abundance relative to control.

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Therefore, linuron and other phenylurea herbicides may act to suppress protein translation.

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Other examples of proteins that were regulated by linuron included phosphoethanolamine methyltransferase that decreased in abundance, and an increased abundance of both DDRGK

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domain-containing protein 1 and peroxiredoxin (Martyniuk et al., 2012). Larger fish have also

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been studied for adverse responses related to linuron exposure. In another transcriptomic study, male brown trout were exposed to three concentrations of linuron (one of either 2.5. 25 and 250

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µg/L) for four days (Uren Webster et al., 2015). Liver cholesterol was significantly reduced, and this was also associated with a reduction in the expression of transcripts related to cholesterol

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biosynthesis. The study also reported that there was evidence for an oxidative stress response in brown trout, as glutathione-related anti-oxidants were increased at the mRNA level (Uren

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Webster et al., 2015). Although there are some data for linuron in terms of molecular responses in fish, there are no data currently in frogs and to the best of our knowledge, no data on the molecular impacts of diuron. To address this knowledge gap, and to augment the information from the aforementioned studies, we leveraged different in silico approaches to learn more about putative modes of action that are related to these two widely studied phenylurea herbicides, diuron and linuron.

6.1 Ureic-based herbicides in ToxCast™ and Tox21: New modes of action in non-target organisms?

A total of 8 ureic based compounds (diuron, fenuron, fluometuron, linuron, monuron, nicosulfuron, tebuthiuron, and thidiazuron) were identified as having been screened during the

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ACCEPTED MANUSCRIPT ToxCast™ program as of 2014. ToxCast™ represents a number of high throughput screening assays to expose living cells or isolated proteins to chemicals and has been leveraged to address

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the mode of action for endocrine disrupting chemicals (Dreier et al., 2015; Rotroff et al., 2013;

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Wills et al., 2015) and mitochondrial inhibitors (Wills et al., 2015). Data are manually curated into those chemicals that activate/inhibit a particular assay (active call) or those that do not

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activate an assay (inactive call). When mining the database, it was discovered that there were a

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total of 3407 bioassays and 254 active hits for ureic-based chemicals in the database. Approximately 75% of the assays receiving active calls were those assays responsive to diuron,

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linuron, and thidiazuron (Figure 2).

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From the ToxCast™ database (release 2014), two bioassays were activated by 5 of the 8

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ureic based compounds (diuron, fluometuron, linuron, monuron, thidiazuron) and these were

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NVS_ENZ_rMAOAC (measures conversion of 14C-Serotonin into 14C-5-hydroxy indoleacetaldehyde (HIA)) and NVS_ADME_hCYP1A2 (target is CYP1A2) (Supplemental Table S1). Thus, two unexplored modes of action of ureic-based herbicides in fish and amphibians may be to act on serotonin production, as well as drug metabolism and the synthesis of cholesterol as CYP1A2 plays a significant role in metabolizing this steroid precursor in mammals. Four of the 8 compounds also activated bioassays that included CYP2A1, the pregnane X receptor which is a steroid and xenobiotic sensing nuclear receptor, and the NRF2/ARE signaling pathway which coordinates xenobiotic metabolism and mobilizes antioxidant defense mechanisms. Thus, the bioassays most highly activated by urea-substituted herbicides were those corresponding to the inhibition of CYP enzymes (1A2 and 2A1), antioxidant response element (ARE), and monoamine oxidase A. We also queried the latest release of Tox21 data (Oct. 2015) and identified some assays that were activated by 23

ACCEPTED MANUSCRIPT buturon, diuron, fenuron, isoproturon, linuron, monolinuron, monuron, nicosulfuron, siduron, tebuthiuron, and thidiazuron. Thidiazuron, linuron, and diuron comprised the majority of active

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calls (~50%) compared to the other urea based herbicides queried. Activity for

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TOX21_AhR_LUC_Agonist, TOX21_Aromatase_Inhibition, and

compounds activate AhR and block AR signaling.

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TOX21_AR_BLA_Antagonist_ratio were noted, supporting the hypothesis that these

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A final point to make is that these assays are based on mammalian cell lines and may

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provide limited insight into responses in non-mammalian species. For example, in the case of CYP1A2, there is 72% sequence identify between the human CYP1A2 for zebrafish, thus it is

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expected that there will be differences in species sensitivity to herbicide exposure and CYP1A2

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may not have the same functions as that described in mammalian systems. Despite this

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limitation, these data provide new direction for research on toxicity pathways and provides a rationale for assessing certain functional endpoints in non-mammalian species. Further details on the assays as well as the AC50 values (concentration at half maximal activity) are provided in Supplemental Table S2.

6.2 Prediction of Activity Spectra of Substances To gain information into the theoretical bioactivities of linuron and diuron based on their structure, we obtained the simplified molecular-input line-entry system (SMILES) format and uploaded these to Prediction of Activity Spectra of Substances (http://www.way2drug.com/PASSOnline) for biological activity prediction (Filimonov et al., 2014). Canonical SMILES formats were the following: diuron (CN(C)C(=O)NC1=CC(=C(C=C1)Cl)Cl) and linuron (CN(C(=O)NC1=CC(=C(C=C1)Cl)Cl)OC). All MOAs and potential interactions with enzymes 24

ACCEPTED MANUSCRIPT identified by the PASS program can be found in Supplemental Data 1. Pa (probability "to be active") estimates the chance that the studied compound is belonging to the sub-class of active

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compounds (resembles the structures of molecules, which are the most typical in a sub-set of

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"actives" in PASS training set) while Pi (probability "to be inactive") estimates the chance that the studied compound is belonging to the sub-class of inactive compounds (resembles the

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structures of molecules, which are the most typical in a sub-set of "inactives" in the PASS

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training set).

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PASS identified 22 putative MOAs in common for both diuron and linuron that included Ubiquinol-cytochrome-c reductase inhibitor, Phospholipid-translocating ATPase inhibitor,

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Cytochrome P450 stimulant, NADPH peroxidase inhibitor, TNF expression inhibitor, L-

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glutamate oxidase inhibitor, Chloride peroxidase inhibitor, and Calcium channel (voltagesensitive) activator. These processes showed a minimum difference of 0.3 for a probability of

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being active compared to inactive against these enzymes. The top ten putative MOAs for diuron and linuron are presented in Table 3. Based upon these data, it is hypothesized that these urea substituted herbicides are most strongly associated with the following MOAs (1) Cytochrome P450 stimulant (2) Ubiquinol-cytochrome-c reductase inhibitor (3) Phospholipid-translocating ATPase inhibitor, and (4) NADPH peroxidase inhibitor. These analyses offer some new endpoints for investigation in toxicological studies based on the chemical structure of these herbicides, for example affects on mitochondrial bioenergetics due to inhibition of ubiquinolcytochrome-c reductase and disruptions in redox balance via inhibition of NADPH peroxidase. 6.3 Gene and pathways altered by diuron and linuron: Leveraging the CTD database

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ACCEPTED MANUSCRIPT Comparative Toxicogenomics Database (CTD) is a robust, publicly available database that aims to advance understanding about how chemicals interact with genes/proteins, and how

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they are associated to disease (Davis et al., 2015; Mattingly et al., 2003). Omics datasets are

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integrated with pathway data (leveraging KEGG and REACTOME pathways) to better characterize the adverse effects of chemicals. The language is a structured hierarchical

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interaction-type vocabulary that describes relationships among genes and proteins that include

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binding and expression, among many other types of interaction. Thus, it is a rich source of

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integration for chemical–gene/protein interactions.

We used the CTD to identify the genes and pathways most likely to be sensitive to diuron

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and linuron (Figure 3). Genes most responsive to diuron included CYP1A1, aryl-hydrocarbon

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receptor (AHR), androgen receptor (AR), aldehyde dehydrogenase 3 family member A1

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(ALDH3A1), and insulin growth factor binding proteins (3 and 6) while linuron was determined to regulate genes such as nuclear receptor subfamily 1, group I, member 2 (Pregnane X Receptor), hydroxy-delta-5-steroid dehydrogenase, 3 beta- and steroid delta-isomerase 1 (3BHSD), and bone morphogenetic protein 4 (BMP4). The effect on transcripts associated to Pregnane X receptor signaling supports in vitro data collected in the ToxCast dataset as Pregnane X receptor activation was observed with 5 out of the 8 compounds. The most prevalent interactions between linuron and targets was that of both AHR and AR (Supplemental Figure 3). This in silico approach supports experimental data reported by Uren Webster et al. (2015) who showed that there was a significant increase in transcripts encoding cyp1a in the liver of brown trout. This suggests that activation of the AHR by linuron is an event that occurs in vivo in fish. Further support for an in vivo interaction between linuron and AHR signaling comes from

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ACCEPTED MANUSCRIPT Ornostay et al. (2013) who demonstrated that linuron increased a gene network that involved aryl

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hydrocarbon receptor nuclear translocator.

Although the numbers of genes regulated by linuron that are collected in the CTD

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database are few, there was substantial gene information for diuron. Interestingly, diuron showed

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high overlap in gene response with those regulated by progesterone, dietary fats, and 17βestradiol (Figure 4). Thus, transcriptomics datasets suggest that these chemicals can elicit

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transcript responses that are similar to those regulated by reproductive hormones. The CTD also

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supports that hypothesis that diuron affects gene targets in a similar way as that observed with diethylnitrosamine, titanium dioxide, phenobarbital, ethinyl-estradiol, copper nanotubes, carbon,

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tamoxifen and vinclozolin (p-value <0.01) (Table 4). Based on data compiled the CTD, diuron

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regulates transcripts that are associated to the following signaling pathways; immune (cytokine-

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cytokine receptor interaction, complement and coagulation cascades), metabolism (drug metabolism - cytochrome P450, glutathione metabolism), Wnt signaling pathway, calcium signaling pathway, and extracellular receptor matrix (ECM)-receptor interaction among others while also affecting genes related to diseases such as cancer (Table 5). Linuron also disrupts transcripts related to prostate cancer, steroid hormone biosynthesis, cancer, and metabolism of xenobiotics by cytochrome P450. The five KEGG pathways in common for both diuron and linuron were “Pathways related to cancer”, “Metabolism of xenobiotics by cytochrome P450”, “Steroid hormone biosynthesis”, “Oocyte meiosis”, and “Prostate cancer”. The effects of these herbicides on endpoints related to reproduction, cell growth and differentiation, and immune system should therefore be examined as potential targets of toxicity.

7. Conclusions and recommendations for future direction

27

ACCEPTED MANUSCRIPT Based on the presence of two widely used phenylurea herbicides, diuron and linuron, in surface waters and ground waters reported in several geographic areas in North America, it

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appears that the environment in this continent creates conditions conducive to their persistence

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and mobility in soil and water. Whether other phenylurea herbicides in use or that may come into use in North America will follow the same trend of low-level ng to µg/L water

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contamination after agricultural applications is unknown. The results of a literature review of

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diuron and linuron were: these are the only two phenylurea herbicides monitored in the environment; most aquatic toxicity studies of these pesticides to date have focussed on adult

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acute toxicity studies in fish with a more recent emphasis on early life stages of fish; and, that toxicity data are deficient for amphibians. With respect to mode of action and sub-lethal adverse

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effects of a phenylurea, linuron is the most well studied in vertebrates and has recently been deemed an anti-androgen both in vitro and in vivo after testing in the USEPA’s Endocrine

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Disruptor Screening Programs Tier 1 assay battery. Although the studies on early life stages of fish to date are few in number and are mainly focused on diuron, they do represent a wide range of experimental designs and test species, and indicate that that adverse sub-lethal effects on development and behaviour are more likely to occur at concentrations that are ~5-10 fold higher than what is presently found in surface waters. However, the co-occurrence of multiple phenylurea herbicides and other contaminants with potential additive, synergistic or antagonistic effects that represents a more likely environmental exposure scenario is a concern, and such contaminant cocktail effects are poorly understood. To obtain a more thorough understanding of the impacts of low level phenylurea contamination in real-world exposure scenarios, future studies examining a wide range of teleost species including various developmental stages and the lethal and sub-lethal effect levels of phenylurea herbicides alone and in combination with other

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ACCEPTED MANUSCRIPT environmental contaminants are necessary. Indeed, the single guideline for a phenylurea herbicide in Canada for the protection of aquatic life is 7 µg/L and the Canadian regulatory

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agency responsible for pesticide registration has recently decided to phase out the use of linuron

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due to human health and environmental risks. Again, while toxicity studies in amphibians are insufficient to draw conclusions, based on the few toxicity studies in this taxa summarized in this

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review, there is some evidence that environmentally relevant levels of diuron and linuron may

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cause adverse effects in early aquatic larval stages of frogs. The effects of phenylurea herbicides on adult, terrestrial/aquatic amphibians as a whole is unknown, and may be just one more

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contributing factor to global amphibian population declines.

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Another knowledge gap is in regard to the mode(s) of action and adverse effects of other

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lesser-studied phenylurea herbicides that are used today, for example thidiazuron. Thidiazuron

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emerged as one of the phenylureas of concern, due to its relatively high activity in many assays in ToxCast™ (putative androgen receptor antagonism, aromatase inhibition), yet data in aquatic organisms are non-existent for this chemical. Studies for several phenylurea herbicides in fish and frogs are needed to determine if there are conserved molecular responses to these herbicides in aquatic vertebrates; indeed many of these compounds remain untested in aquatic wildlife. Of interest, the widely publicized and controversial herbicide atrazine, as well as cyanazine and simazine, are also inhibitors of photosynthesis at photosystem II, and there are widespread reports of endocrine disruption in fish and amphibians from a chemical like atrazine at low doses. Phenylurea herbicides are inhibitors of photosynthesis at photosystem II Site B (as oppose to Site A inhibition by atrazine) and linuron has already been shown to be anti-androgenic. Therefore, one hypothesis to be tested is whether or not phenylurea herbicides have similar endocrine disrupting effects in vertebrates since they target the same complex in plants. 29

ACCEPTED MANUSCRIPT To begin to address the data gaps for phenylurea herbicides and their mode of action, we synthesized in silico information based upon (1) receptor binding and DNA binding assays

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(ToxCast™) (2) gene expression endpoints (CTD) and (3) chemical structure-function

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relationships to help guide future investigations into adverse effects of these compounds in vertebrate models. High throughput data gathered from ToxCast™ revealed that urea-substituted

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herbicides activated bioassays related to inhibition of androgen receptors in addition to CYP

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enzymes (1A2 and 2A1), antioxidant response element (ARE), and monoamine oxidase A. Transcriptomic data collected from the Comparative Toxicogenomics Database suggested that

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diuron mimics the action of hormones (progesterone, ethinyl estradiol), metals (titanium dioxide, copper), pharmaceutics (phenobarbital), other anti-androgenic compounds such as the fungicide

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vinclozolin, and carbon nanotubes. Prediction of Activity Spectra of Substances revealed that diuron and linuron are significantly associated with the following MOAs: (1) Cytochrome P450

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stimulant (2) Ubiquinol-cytochrome-c reductase inhibitor and (3) Phospholipid-translocating ATPase inhibitor.

The next question was whether the different in silico approaches converged on any common mechanisms and whether this was consistent with experimental data. Based upon the weight of evidence from the in silico analysis from both ToxCast™ and the CTD database, Pregnane X Receptor appears to be a likely target for ureic-based herbicides, in addition to CYP enzymes (e.g. PASS identified that both diuron and linuron were Cytochrome P450 stimulants). Activation of Cyp1A2 assays in ToxCast, transcriptomic profiles from CTD and KEGG (“Metabolism of xenobiotics by cytochrome P450”, “Steroid hormone biosynthesis”) and experimental data from Ornostay et al (2013) and Uren Webster et al. (2015) suggest that steroid biosynthesis and cholesterol metabolism are also common targets for ureic-based herbicides. 30

ACCEPTED MANUSCRIPT Collectively, we also suggest future investigations should quantify the effects of these herbicides on endpoints related to cell growth and differentiation, and the immune system as these may be

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potential targets of toxicity based on different in silico analyses.

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As a final note, it is important to point out the strengths and limitations for in silico

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approaches such as these to predict MOAs. While the chemical spaces covered in the ToxCast and Tox21 programs are impressive and relatively extensive (1000s of chemicals), the assays

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implemented are mammalian-based gene assays in model species such as rat, mouse, and human.

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Receptor-ligand interaction and gene activation can be species specific, and ligand-binding kinetics will likely proceed differently in fish and amphibians. Species-specific responses may be

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one reason why models using these in vitro data perform better in predicting human toxicity end

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points rather than animal toxicity (Huang et al., 2016). Moreover, another challenge with the

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cell-based assays in ToxCast and Tox21 is cytotoxicity; a chemical may be classified as active in some cases for an assay while exhibiting a cytotoxic response, leading to false positives/negatives. In terms of the Omics approaches, the user-friendly Comparative Toxicogenomics Database is designed to integrate chemical-gene/protein interactions and chemical- and gene/protein-disease relationships by accessing 23 external databases. This is a strong approach to integrate such complex data. However, a disadvantage of the transcriptomic data is that there can be significant data gaps; not every chemical has been tested in every relevant species. The lack of standardization and variability in transcriptome responses can also be a challenge (Dreier et al., 2016). However, RNA-seq approaches are anticipated to fill in some of these data gaps as more researchers leverage these types of tools. Despite these limitations, high throughput approaches offer a wealth of information that can be carefully mined and synthesized, and the aim here was to leverage some of these tools, including experimental 31

ACCEPTED MANUSCRIPT data in fish and frogs, to convergence on target pathways for ureic-based herbicides (weight of evidence approach). Comparative ecotoxicology is anticipated to grow, and efforts in the future

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should be directed towards developing similar programs for ecologically relevant species for

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monitoring environmental pollutants.

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Acknowledgements

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The authors have no conflict of interest. The authors thank the University of Florida College Of Veterinary Medicine (CJM) and the Department of Biological Sciences of Simon Fraser University (VLM) for funding.

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References

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Davy, M., Shaugnessy, W.J., 2008. Risks of linuron use to federally threatened California redlegged frog (Rana aurora draytonii), in: O.P.P. Environmental Fate and Effects Division, US EPA, Washington, D.C. 20460 (Ed.). Dreier, D.A., Connors, K.A., Brooks, B.W., 2015. Comparative endpoint sensitivity of in vitro estrogen agonist assays. Reg. Toxicol. Pharmacol., 72: 185-193. Dreier, D.A., Loughery, J.R., Denslow, N.D., Martyniuk, C.J., 2016. The influence of breeding strategy, reproductive stage, and tissue type on transcript variability in fish. Comparative biochemistry and physiology. Part D, Genomics & Proteomics. 19: 151-158. Du Pont Canada, I., 1973. Acute toxicity of H-7952, MR-581 to bluegill (Lepomis macrochirus) and rainbow trout (Salmo gairdneri). INZ-326. Du Pont Canada Inc., Mississauga, ON. Du Pont Canada, I., 1986. Static acute 96-hour LC50 of linuron (INZ-326-118) to rainbow trout (Salmo gairdneri). HLR no.525-86, MR no 4581-420. Du Pont Canada Inc., Mississauga, ON. Du Pont Canada, 1989. DuPont Corporation data. In: [PPD] Agricultural Research Service Pesticide Properties Database [Internet]. 2009. Beltsville (MD): Crop Systems and Global Change Laboratory in Beltsville. [cited 2009 Nov 19]. Available from: http://www.ars.usda.gov/Services/docs.htm?docid=14199. Environment Canada, 2011. Presence and levels of priority pesticides in selected Canadian aquatic ecosystems, ISBN 978-1-100-18386-2 in: Water Science and Technology Directorate (Ed.)., Environment Canada. Francis, B.M., Lampman, R.L., Metcalf, R.L., 1985. Model ecosystem studies of the environmental fate of five herbicides used in conservation tillage. Arch. Environmen. Cont. Toxicol. 14: 693-704. Frank, R., Ripley, B.D., Braun, H.E., Clegg, B.S., Johnson, R., O'Neill, T.J., 1987. Survey of farm wells for pesticides residues, Southern Ontario, Canada, 1981–1982, 1984. Arch. Environmen. Cont. Toxicol. 16: 1. Gagnon, M.M., Rawson, C.A., 2009. Diuron increases spinal deformity in early-life-stage pink snapper Pagrus auratus. Marine Poll. Bull. 58: 1083-1085. Gatidou, G., Thomaidis, N.S., Zhou, J.L., 2007. Fate of Irgarol 1051, diuron and their main metabolites in two UK marine systems after restrictions in antifouling paints. Environment International 33. Greulich, K., Hoque, E., Pflugmacher, S., 2002. Uptake, metabolism, and effects on detoxication enzymes of isoproturon in spawn and tadpoles of amphibians. Ecotoxicol. Environ. Saf. 52: 256-266. Hausmann, S.M., 1992. Anaerobic aquatic metabolism of [phenyl(U)-14C]diuron. Laboratory Project ID: AMR 2067-91. Unpublished study performed and submitted by E.I. du Pont de Nemours and Company, Wilmington, DE. Hausmann, S.M., Kraut, G.M., 1992. Aerobic aquatic metabolism of [phenyl(U)14C]diuron.Laboratory Project ID: AMR 2066-91. Unpublished study performed and submitted by E. I. duPont de Nemours and Company, Wilmington, DE. Hawkins, D.R., Kirkpatrick, D., Shaw, D., Chan, S.C., 1990. The metabolism of [phenyl(U)14C]diuron in Keyport silt loam soil under aerobic conditions. Du Pont Report No. AMR1202-88. Huntingdon Research Center Report No. HRC/DPT 189/891860. Unpublished study performed by Huntingdon Research Centre, Huntingdon, Cambridgeshire, England, and submitted by E.I du Pont de Nemours & Company, Inc., Wilmington, DE.

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Rotroff, D.M., Dix, D.J., Houck, K.A., Knudsen, T.B., Martin, M.T., McLaurin, K.W., Reif, D.M., Crofton, K.M., Singh, A.V., Xia, M., Huang, R., Judson, R.S., 2013. Using in vitro high throughput screening assays to identify potential endocrine-disrupting chemicals. Environ. Health Perspect. 121: 7-14. Saglio, P., Trijasse, S., 1998. Behavioral responses to atrazine and diuron in goldfish. Arch. Environ. Cont. Toxicol. 35: 484-491. Schneiders, G., 1990. Aerobic Soil Metabolism of Phenyl(U)- Carbon 14 Linuron in Hanford Sandy Loam: Lab Project Number: AMR-134888. Unpublished study prepared by E. I. du Pont de Nemours and Co. 47 p. Schuler, L.J., Rand, G.M., 2008. Aquatic risk assessment of herbicides in freshwater ecosystems of South Florida. Arch. Environ. Contam. Toxicol. 54: 571-583. Schuytema, G.S., Nebeker, A.V., 1998. Comparative toxicity of diuron on survival and growth of Pacific treefrog, bullfrog, red-legged frog, and African clawed frog embryos and tadpoles. Arch. Environ. Contam. Toxicol. 34: 370-376. Stephenson, R.R., and D.F. Kane. , 1984. Persistence and effects of chemicals in small enclosures in ponds. Arch. Environ. Cont. Toxicol. 13: 313–326. Stidham, M.A., 1991. Herbicides that inhibit acetohydroxyacid synthase. Weed Science, 428434. Tomlin, 2005-2006. The e-pesticide manual [CD-ROM], in: T. CDS (Ed.), 13th ed. British Crop Production Council, Alton (UK). Tucker, C.S., Kingsbury, S.K., Ingram, R.L., 2003. Tissue residues of diuron in channel catfish Ictalurus punctatus exposed to the algicide in consecutive years. J. World Aquacult. Soc. 34: 203-209. Turner, K.J., McIntyre, B.S., Phillips, S.L., Barlow, N.J., Bowman, C.J., Foster, P.M., 2003. Altered gene expression during rat Wolffian duct development in response to in utero exposure to the antiandrogen linuron. Toxicol. Sci. 74: 114-128. Uren Webster, T.M., Perry, M.H., Santos, E.M., 2015. The Herbicide Linuron Inhibits Cholesterol Biosynthesis and Induces Cellular Stress Responses in Brown Trout. Environ. Sci. Technol. 49: 3110-3118. USEPA, 1992. Pesticides in Ground Water Database- A compilation of Monitoring Studies: 1971 - 1991. Office of Prevention, Pesticides, and Toxic Substances, EPA 734-12-92-001. USEPA, 1995. Reregistration Eligibility Decision (RED): Linuron. U.S. Environmental Protection Agency (EPA), Washington, DC. March, 1995. EPA 738-R-95-003. USEPA, 2003. Reregistration eligibility decision for diuron, List A, Case 0046, in: USEPA (Ed.). USEPA, 2010. Linuron Final Work Plan, Registration Review, December 2010, Case Number 0047, EPA-HQ-OPP-2010-0228, Pesticide Re-evaluation Divsion, US Environmental Protection Agency. USEPA, 2015. EDSP: Weight of evidence analysis of potential interaction with the estrogen androgen or thyroid pathways, chemical: linuron. Office of Pesticide Programs, Office of Science Coordination and Policy. USGS, 1998. National Water Quality Assessment (NAWQA), Pesticides National Synthesis Project. Online at http://ca.water.usgs.gov/pnsp/allsum/ - over and http://ca.water.usgs.gov/pnsp/streamsum/streamT1.html.

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Van Rensen, J.J.S., 1989. Herbicides interacting with photosystem II. Herbicides and Plant Metabolism, Society for Experimental Biology Seminar Series 38, A.D. Dodge (Ed.), Cambridge University Press, New York, New York, USA. p. 21-25. Wills, L.P., Beeson, G.C., Hoover, D.B., Schnellmann, R.G., Beeson, C.C., 2015. Assessment of ToxCast Phase II for Mitochondrial Liabilities Using a High-Throughput Respirometric Assay. Toxicol. Sci. 146: 226-234. Wolf, C., Lambright, C., Mann, P., Price, M., Cooper, R.L., Ostby, J., Gray, L.E., 1999. Administration of potentially antiandrogenic pesticides (procymidone, linuron, iprodione, chlozolinate, p, p'-DDE, and ketoconazole) and toxic substances (dibutyl-and diethylhexyl phthalate, PCB 169, and ethane dimethane sulphonate) during sexual differentiation produces diverse profiles of reproductive malformations in the male rat. Toxicol. Industrial Health 15: 94-118. Woudneh, M.B., Ou, Z., Sekela, M., Tuominen, T., Gledhill, M., 2009. Pesticide multiresidues in waters of the Lower Fraser Valley, British Columbia, Canada. Part I. Surface water. Journal Environm. Quality 38: 940-947. Xing, Z., Chow, L., Cook, A., Benoy, G., Rees, H., Ernst, B., Meng, F., Li, S., Zha, T., Murphy, C., Batchelor, S., Hewitt, L.M., 2012. Pesticide application and detection in variable agricultural intensity watersheds and their river systems in the maritime region of Canada. Arch. Environ. Cont. Toxicol. 63: 471-483. Yu, W.C., 1988. Anaerobic soil metabolism of [phenyl(U)-14C]diuron, 30 August 1988. Cambridge Analytical Associates.

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Figure 1. Examples of some phenyl-urea based chemicals, depicting the wide array of structures and substitutions. Structures were compiled from the EMBL-Chemical Entities of Biological Interest (ChEBI). Research has primarily focused on diuron and linuron for aquatic species.

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Figure 2. Percent total activation of ToxCast bioassays for the eight ureic-based compounds tested in the program. Thidiazuron and linuron comprised >50% of the positive or active “hits” in high throughput screening assays.

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Figure 3. The comparative toxicogenomics database reveals that (A) linuron and (B) diuron regulate both androgen receptor and the aryl hydrocarbon receptor. The top 10 transcripts are presented as well as the number of interactions identified in the database. The X-axis represents the number of studies that have identified that the transcript is modulated in expression by either linuron or diuron.

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Figure 4. Progesterone, dietary fats, and 17β-estradiol show related gene set responses to diuron.

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Table 1: Summary of selected physical/chemical properties and persistence in water and soil of diuron and linuron Reference Diuron

Linuron

Molecular weight (g/mol) Log Kow (octanol–water partition coefficient)

233.1

249.1

Tomlin, 2005-2006 Tomlin, 2005-2006

2.85 ± 0.03

3.00

Tomlin, 2005-2006

2.4 ± 0.2

3.28 ± 0.15

Thomas et al., 2002

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81

1.1 x 10-7

2.0 x 10-3

Tomlin, 2005-2006

7.04 x 10 -6

6.1 x 10-8

Tomlin, 2005-2006

Jewett & Koper, 2016

41 (silt loam sediment:water system) 15 (sand sediment:water system)

Hausmann & Kraut, 1992; USEPA, 2003

Jewett & Koper, 2016

Henry’s Law constant, caclulated (Pa·m3 /mol)

33 (clay loam sediment:water system)

Anaerobic aquatic metabolism half-life at 2425 °C (days)

5 (clay loam sediment:water system)

Persistence in soil Aerobic soil metabolism half-life at 24-25 °C (days)

372 (silt loam)

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Persistence in water Aerobic aquatic metabolism half-life at 2425 °C (days)

Anaerobic soil metabolism half-life at 24-25 °C (days)

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Water solubility (mg/L at 25 °C) Vapour pressure (Pa) (Pa at 25 °C)

MA

Log Koc (organic carbon–water partition coefficient) (dimensionless)

Diuron

1000 (silt loam)

<21 silt loam and Hausmann, 1992; sand:water systems USEPA, 2003

52 (sandy loam) 1100 (loam) 171 (sandy loam) 202 (sandy clay loam) 127(sandy loam) nd

Linuron

PT

Parameter

Kidd & James, 1991

Kidd & James, 1991 Kidd & James, 1991 Kidd & James, 1991

Monson, 1986; USEPA 1995

Hawkins et al., Schneiders, 1990; 1990; USEPA 2003 USEPA 1995

Yu, 1988; USEPA 2003

nd

nd, refers to no data

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Linuron mg/L

Diuron mg/L

Goldfish (Carrasius auratus)

juvenile

behavior

0.005

1

Goldfish

juvenile

behavior

0.05

1.5

Pink snapper (Pagnus auratus) embryo, larva

development

2

Turbot (Psetta maxima)

embryo

survival

4

Turbot

larva

survival

4

Fathead minnow (Pimphales promela)

larva

xenobiotic metabolism

4

Fathead minnow

larva

survival

1, 2, 4, 8

Fathead minnow

juvenile

7

Fathead minnow

embryo-larva

7

Fathead minnow

embryo-larva

10

Fathead minnow

juvenile

10

Fathead minnow

juvenile

15*

Golden medaka (Oryzias latipes)

15*

Golden medaka (Oryzias latipes)

35

Rainbow trout (Oncorhynchus fingerlings mykiss)

60

Fathead minnow

Saglio and Trijasse, 1998; decreased grouping behavior at 0.005 but not 0.05 mg/L

Saglio and Trijasse, 1998; increase of burst swimming reactions defined as sudden spurt of nondirected movement, followed by immobilization of the fish

0.05

Gagnon et al. 2009; decrease in proportion of eggs hatched and developed normally up to 36 h and increased spinal deformity rate in hatched larvae

10.076

Lazhar et al. 2012; concentration causing 50% mortality (LC50) during embryogenesis

7.826

Lazhar et al. 2012: concentration causing 50% mortality (LC50) for early larval stage

2.797

Newman et al. 2001; 50% inhibition of soluble epoxide hydrolase activity in whole body protein extracts after in vivo exposure

7.458

Newman et al. 2001; concentration causing 70% average mortality

survival

23.3, 19.9, 14.2, 7.7

Call et al. 1987; LC50 after exposure durations of 24, 48, 96, 192 hours, respectively

growth

8.3, 15.1, 31.2

Nebeker et al. 1998; decreased total body length (developmental endpoints not reported)

survival

11.7

Nebeker et al. 1998; LC50 for embryo/larva

growth

3.4, 6.5, 12.2, 20.0, 27.1

Nebeker et al 1998; decreased total body length and weight

survival

27.1

Nebeker et al. 1998; LC50 for juvenile

embryo-larva

morphological abnormalities

7.459, 15.151

Newman et al. 2001; 100% displayed failed swim bladder inflation and glall bladder enlargement, 90% displayed yolk sac edema

embryo-larva

development

2.005, 3.963

Newman et al. 2001; delayed time to hatch

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Source and detailed effect

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Table 2: Effects of linuron and diuron on early life stages of teleosts Duration (d) Species Developmental Effect on: Stage

embryo-larvajuvenile

renal and liver histopathology development & survival

0.03, 0.12,0.24

Oulmi et al. 1995; non-lethal ultrastructural alterations in liver and kidney in dose dependent manner 0.078

Call et al. 1987; increased gross deformities and decreased survival

*plus 4-5 day depuration

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Table 3: The theoretical bioactivities of linuron and diuron were estimated through Prediction of Activity Spectra of Substances (PASS).

0.779 0.766 0.765 0.714 0.713 0.696

0.935 0.907 0.857 0.648 0.595 0.557 0.562 0.532 0.536 0.546

0.001 0.005 0.015 0.021 0.014 0.02 0.035 0.013 0.03 0.057

0.934 0.902 0.842 0.627 0.581 0.537 0.527 0.519 0.506 0.489

Cytochrome P450 stimulant Ubiquinol-cytochrome-c reductase inhibitor Phobic disorders treatment Phospholipid-translocating ATPase inhibitor Analgesic, non-opioid Antiallergic Antiarthritic Cell adhesion molecule inhibitor Analgesic NADPH peroxidase inhibitor

PT ED

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0.004 0.015 0.004 0.005 0.005 0.047

CE

Linuron

0.783 0.781 0.769 0.719 0.718 0.743

Activity Ubiquinol-cytochrome-c reductase inhibitor Phospholipid-translocating ATPase inhibitor Phobic disorders treatment Glycosylphosphatidylinositol phospholipase D inhibitor Cytochrome P450 stimulant NADPH peroxidase inhibitor N-acylmannosamine kinase inhibitor TNF expression inhibitor Eye irritation, inactive Membrane integrity agonist

AC

Herbicide Pa Pi delta Diuron 0.964 0.002 0.962 0.89 0.003 0.887 0.882 0.009 0.873 0.797 0.014 0.783

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Similarity Index 0.148856 0.144576 0.138539 0.13822 0.134799 0.133204 0.13201 0.131616 0.131153 0.128001 0.127223 0.126785 0.126452 0.12617 0.124391 0.123249 0.123202 0.121769 0.121662 0.121448

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Progesterone Dietary Fats Estradiol Carbon Tetrachloride Diethylnitrosamine Titanium dioxide Cobaltous chloride Phenobarbital Dexamethasone Ethinyl Estradiol Ammonium Chloride Dibutyl Phthalate Copper Nanotubes, Carbon Tretinoin Palm oil Propylthiouracil Pirinixic acid Tamoxifen Vinclozolin

Chemical CAS RN ID D011374 57-83-0 D004041 N/A D004958 50-28-2 D002251 56-23-5 D004052 55-18-5 C009495 13463-67-7 C018021 7646-79-9 D010634 1950-06-06 D003907 1950-02-02 D004997 57-63-6 D000643 12125-02-9 D003993 84-74-2 D003300 7440-50-8 D037742 N/A D014212 302-79-4 C041786 8002-75-3 D011441 51-52-5 C006253 50892-23-4 D013629 10540-29-1 C025643 50471-44-8

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Chemical/Hormones

PT

Table 4: Chemicals and hormones that show significant similarity to diuron in the transcripts activated (P<0.01)

46

ACCEPTED MANUSCRIPT

KEGG:01100

Focal adhesion

KEGG:04510

1.03E-32

Pathways in cancer

KEGG:05200

4.84E-20

Regulation of actin cytoskeleton Cytokine-cytokine receptor interaction Endocytosis

KEGG:04810

5.74E-16

KEGG:04060

6.01E-12

35

KEGG:04144

2.13E-13

33

ECM-receptor interaction

KEGG:04512

1.21E-25

32

Ribosome

KEGG:03010

PT ED

Annotated Genes Quantity 170

9.95E-22

30

Amoebiasis

KEGG:05146

4.73E-20

30

MAPK signaling pathway

KEGG:04010

3.65E-08

30

Protein processing in endoplasmic reticulum Complement and coagulation cascades Drug metabolism cytochrome P450 Chemokine signaling pathway Metabolism of xenobiotics

KEGG:04141

1.28E-11

28

KEGG:04610

6.48E-21

26

KEGG:00982

7.67E-20

26

KEGG:04062

3.01E-09

26

KEGG:00980

3.19E-19

25

SC

RI

Metabolic pathways

Corrected Pvalue 3.65E-71

NU

52

MA

CE

Diuron

Pathway ID

AC

Chemical Pathway

PT

Table 5. KEGG pathways significantly affected by diuron and linuron. Only those KEGG pathways regulated by diuron that contained more than 20 members in the pathway are shown.

49 36

Genome Frequency 1216/37558 genes: 3.24% 211/37558 genes: 0.56% 333/37558 genes: 0.89% 220/37558 genes: 0.59% 277/37558 genes: 0.74% 220/37558 genes: 0.59% 88/37558 genes: 0.23% 96/37558 genes: 0.26% 108/37558 genes: 0.29% 283/37558 genes: 0.75% 180/37558 genes: 0.48% 70/37558 genes: 0.19% 76/37558 genes: 0.20% 193/37558 genes: 0.51% 72/37558 47

ACCEPTED MANUSCRIPT

Linuron

1.36E-09

25

Protein digestion and absorption Leukocyte transendothelial migration Vascular smooth muscle contraction Tight junction

KEGG:04974

2.76E-16

24

KEGG:04670

2.13E-12

24

KEGG:04270

4.78E-11

KEGG:04530

1.28E-10

Tuberculosis

KEGG:05152

4.61E-08

Toxoplasmosis

KEGG:05145

3.85E-10

PPAR signaling pathway

KEGG:03320

Glutathione metabolism

KEGG:00480

Fc gamma R-mediated phagocytosis Wnt signaling pathway

KEGG:04666

SC

RI

KEGG:05010

PT ED

MA

NU

24 24 24 23 22

3.5E-17

21

2.85E-11

21

KEGG:04310

0.000000579

21

Calcium signaling pathway

KEGG:04020

0.0000043

21

Huntington's disease

KEGG:05016

0.00000764

21

Systemic lupus erythematosus Phagosome

KEGG:05322

0.000000217

20

KEGG:04145

0.00000606

20

Prostate cancer

KEGG:05215

2.80E-12

7

AC

CE

3.94E-15

genes: 0.19% 172/37558 genes: 0.46% 83/37558 genes: 0.22% 119/37558 genes: 0.32% 136/37558 genes: 0.36% 142/37558 genes: 0.38% 186/37558 genes: 0.50% 136/37558 genes: 0.36% 74/37558 genes: 0.20% 54/37558 genes: 0.14% 98/37558 genes: 0.26% 162/37558 genes: 0.43% 181/37558 genes: 0.48% 187/37558 genes: 0.50% 139/37558 genes: 0.37% 168/37558 genes: 0.45%

PT

by cytochrome P450 Alzheimer's disease

92/37558 48

ACCEPTED MANUSCRIPT

8.67E-09

5

KEGG:05200

2.49E-08

7

Glioma

KEGG:05214

3.80E-04

3

Melanoma

KEGG:05218

4.96E-04

Metabolism of xenobiotics by cytochrome P450 Oocyte meiosis

KEGG:00980

5.17E-04

KEGG:04114

0.00211

Dorso-ventral axis formation

KEGG:04320

0.00707

SC

MA

NU

3 3 3 2

AC

CE

PT ED

PT

KEGG:00140

RI

Steroid hormone biosynthesis Pathways in cancer

genes: 0.24% 58/37558 genes: 0.15% 333/37558 genes: 0.89% 65/37558 genes: 0.17% 71/37558 genes: 0.19% 72/37558 genes: 0.19% 115/37558 genes: 0.31% 29/37558 genes: 0.08%

49