- Email: [email protected]
An interspecies correlation model to predict acute dermal toxicity of plant protection products to terrestrial life stages of amphibians using fish acute toxicity and bioconcentration data
Accepted Manuscript An interspecies correlation model to predict acute dermal toxicity of plant protection products to terrestrial life stages of amphibians using fish acute toxicity and bioconcentration data
Lennart Weltje, Philipp Janz, Peter Sowig PII:
S0045-6535(17)31462-5
DOI:
10.1016/j.chemosphere.2017.09.047
Reference:
CHEM 19918
To appear in:
Chemosphere
Received Date:
23 June 2017
Revised Date:
10 September 2017
Accepted Date:
11 September 2017
Please cite this article as: Lennart Weltje, Philipp Janz, Peter Sowig, An interspecies correlation model to predict acute dermal toxicity of plant protection products to terrestrial life stages of amphibians using fish acute toxicity and bioconcentration data, Chemosphere (2017), doi: 10.1016 /j.chemosphere.2017.09.047
This is a PDF file of an unedited manuscript that has been accepted for publication. As a service to our customers we are providing this early version of the manuscript. The manuscript will undergo copyediting, typesetting, and review of the resulting proof before it is published in its final form. Please note that during the production process errors may be discovered which could affect the content, and all legal disclaimers that apply to the journal pertain.
ACCEPTED MANUSCRIPT Highlights for the manuscript: AN INTERSPECIES CORRELATION MODEL TO PREDICT ACUTE DERMAL TOXICITY OF PLANT PROTECTION PRODUCTS TO TERRESTRIAL LIFE STAGES OF AMPHIBIANS Lennart Weltje, Philipp Janz and Peter Sowig
1. 2. 3. 4.
An equation is derived to predict acute dermal toxicity to terrestrial amphibians Fish bioconcentration and acute toxicity data are required as input The model can be used in a screening approach for pesticide risk assessment Application would significantly reduce amphibian (vertebrate) testing
ACCEPTED MANUSCRIPT
1
AN INTERSPECIES CORRELATION MODEL TO PREDICT ACUTE DERMAL
2
TOXICITY OF PLANT PROTECTION PRODUCTS TO TERRESTRIAL LIFE STAGES
3
OF AMPHIBIANS USING FISH ACUTE TOXICITY AND BIOCONCENTRATION
4
DATA
5 6
Lennart Weltjea*, Philipp Janza, and Peter Sowigb
7 8
a
9
Germany
BASF SE, Crop Protection – Ecotoxicology, Speyerer-Strasse 2, D-67117 Limburgerhof,
10
b
11
* author for correspondence. E-mail: [email protected], Tel.: +49 – 62160 - 28564
Bayer CropScience AG, Industriepark Höchst, D-65926 Frankfurt-Höchst, Germany
1
ACCEPTED MANUSCRIPT
12
ABSTRACT
13
This paper presents a model to predict acute dermal toxicity of plant protection products (PPPs) to
14
terrestrial amphibian life stages from (regulatory) fish data. By combining existing concepts,
15
including interspecies correlation estimation (ICE), allometric relations, lethal body burden (LBB)
16
and bioconcentration modelling, an equation was derived that predicts the amphibian median lethal
17
dermal dose (LD50) from standard acute toxicity values (96-h LC50) for fish and bioconcentration
18
factors (BCF) in fish. Where possible, fish BCF values were corrected to 5% lipid, and to parent
19
compound. Then, BCF values were adjusted to an exposure duration of 96 hours, in case steady
20
state took longer to be achieved. The derived correlation equation is based on 32 LD50 values from
21
acute dermal toxicity experiments with 15 different species of anuran amphibians, comprising 15
22
different PPPs. The developed ICE model can be used in a screening approach to estimate the
23
acute risk to amphibian terrestrial life stages from dermal exposures to PPPs with organic active
24
substances. This has the potential to reduce unnecessary testing of vertebrates.
25 26 27 28
Key words: terrestrial toxicity, amphibians, dermal exposure, fish toxicity, risk assessment, lethal
29
body burden; alternative method
2
ACCEPTED MANUSCRIPT
30
1. INTRODUCTION
31
The European Regulation 1107/2009 concerning the placing of plant protection products (PPPs)
32
on the market (EC, 2009) requires the risks for amphibians to be assessed based on available
33
information. Similar wording can be found in the European Food and Safety Authority (EFSA)
34
Guidance Document for risk assessment for aquatic organisms in edge-of-the-field waters (EFSA,
35
2013). Currently, amphibians are not explicitly assessed as it is assumed that the risk assessment
36
for other vertebrates covers the risk to amphibians (both aquatic and terrestrial life stages).
37
However, concerns have recently been voiced regarding potential risks of PPPs to amphibian
38
terrestrial life stages that might not be covered by the current risk assessment (e.g. Belden et al.,
39
2010; Brühl et al., 2011). Against this background, the EFSA has been given a mandate to develop
40
a guidance document for the risk assessment of PPPs for amphibians (and reptiles [note that this
41
paper focuses on amphibians only]). This is a challenging task, given that there are no standard
42
methods available for describing and quantifying amphibian exposure and toxicity. In addition,
43
the life-cycle of amphibians requires consideration of both the aquatic and terrestrial life stages.
44
The situation is further complicated by the demand to minimize testing of vertebrates due to animal
45
welfare (EC Regulation 1107/2009). Therefore, as a first approach it makes sense to consider if
46
the current data requirements on aquatic and terrestrial vertebrates yield data that can be used as
47
surrogates to inform on the sensitivity of amphibians to PPPs.
48
While comprehensive data reviews have shown that the acute and chronic sensitivities of
49
aquatic life stages of amphibians are similar to or lower than those of fish (Aldrich, 2009; Weltje
50
et al., 2013) and thus can be covered in the risk assessment, the situation is less clear for the
51
terrestrial life stages. In general, there are fewer data available on terrestrial than on aquatic life
52
stages. Nevertheless, a recent data comparison of oral acute toxicity values for 26 chemicals
3
ACCEPTED MANUSCRIPT
53
showed that the sensitivity of amphibian terrestrial life stages can be covered by available standard
54
toxicity data on birds and mammals (Crane et al., 2016). However, the exposure route that is
55
probably of highest relevance for terrestrial life stages of amphibians is the dermal exposure route
56
(Fryday and Thompson, 2012; Smith et al., 2007). In contrast to other terrestrial vertebrates,
57
amphibians do not have fur and feathers that might serve as a protective barrier against chemical
58
exposure (Smith et al., 2007). Furthermore, amphibians have a more permeable skin (Quaranta et
59
al., 2009) fulfilling physiological functions such as water and gas exchange (i.e. breathing). In
60
situations where exposure to a PPPs could occur, for instance during a spraying application in a
61
cropped field, this permeability may lead to dermal uptake that is typically not experienced by
62
other vertebrate taxa (Smith et al., 2007). Because of the different properties of the skin,
63
mammalian and avian species are probably not suitable surrogates for dermal exposure routes in
64
the amphibian risk assessment. Accordingly, Berger et al. (2015) found no statistically significant
65
correlation between the acute toxicity of PPPs to amphibians after dermal exposure and the
66
potential of PPPs for skin irritation in rabbits. Therefore, the work presented here aimed at
67
developing a method for predicting acute dermal toxicity in terrestrial amphibian life stages, based
68
on other available data. Such a non-testing method is urgently needed for use in a screening
69
approach to distinguish PPPs of lower from those of higher potential concern. Only when a
70
potential concern is identified, a PPP requires further consideration, which may eventually involve
71
in vivo vertebrate toxicity testing using amphibians. A non-testing approach would thus improve
72
current procedures for risk assessment and avoid unnecessary testing of vertebrates.
73
To derive an equation for acute dermal toxicity estimation in terrestrial life stages of
74
amphibians, some assumptions were made and various concepts established in (aquatic)
75
ecotoxicology were combined. In short, the high correlation between sensitivity of fish to
4
ACCEPTED MANUSCRIPT
76
chemicals and that of amphibian aquatic life stages (Weltje et al., 2013) was assumed to translate
77
to the terrestrial life stage. Further, the lethal body burden (LBB) concept was used to convert LC50
78
values for fish, by multiplication with the bioconcentration factor (BCF), to LD50 values. For
79
dermally exposed amphibians, the LD50 values were calculated by considering the exposed skin
80
surface, body weight, and the exposure rate. The dermally received dose was assumed to be
81
absorbed completely by the exposed amphibians. More detailed explanations are provided in the
82
‘Materials and Methods’ section.
83 84 85
2. MATERIALS AND METHODS
86
2.1 Acute dermal toxicity data for amphibians
87
Acute dermal toxicity data for amphibian terrestrial life stages were collected from Fryday and
88
Thompson (2012) and targeted searches of the scientific literature via Web of Science
89
(http://webofknowledge.com/WOS accessed May 12, 2017) using (amphibian* OR frog* OR
90
toad* OR anura* OR urodel* OR caudat*) AND (dermal OR terrestrial OR overspray). Dermal
91
toxicity data for terrestrial life stages of amphibians are generated mainly for three purposes: i) to
92
study side-effects of chemicals in ecotoxicology (e.g. Vinson et al., 1963); ii) to study chemical
93
means to control invasive amphibian species for animal conservation purposes (e.g. Witmer et al.,
94
2015); iii) to study side-effects of veterinary medicines if applied percutaneously (an alternative
95
to oral application, e.g. Llewelyn et al., 2015). Only studies with dermal exposure routes (i.e.
96
overspray, filter paper, direct application) employing at least 2 dose groups and reporting mortality
97
as an endpoint were included. When formulated PPPs were tested, the application rates were
98
converted to active substance (a.s.). Taylor et al. (1999), Belden et al. (2010) and Brühl et al.
5
ACCEPTED MANUSCRIPT
99
(2013) did not report 50% lethal exposure levels. To obtain these values, they were calculated for
100
each of the tested chemicals from the original data in consideration of: i) the observed mortality at
101
each treatment, ii) any observed control mortality, and iii) the high spacing factor of 10 used in
102
these studies. If more than one 50% lethal effect value for a species was available (e.g. Ferguson
103
and Gilbert, 1967), the geometric mean of these values was used.
104
For converting dermal exposure values to body burdens, the exposed body surface and the
105
body weight of the experimental animals are needed. Therefore, allometric equations were applied
106
for relating exposed skin area to body length and body weight in anurans. If body weight (bw) or
107
body length (i.e. snout to vent length (SVL)) were not reported, SVL (in mm) was estimated based
108
on bw (in g) or vice versa, using the following allometric equation:
109
log[bw] = 2.97 ∙ log[SVL] - 3.96
110
The numerical parameter values were derived from allometric equations for yellow-bellied toad
111
Bombina variegata (Bancila et al. 2010), common toad Bufo bufo (Kuhn 1994), and common frog
112
Rana temporaria (Gibbons and McCarthy 1984; Grözinger et al. 2014) – see Table 1. These
113
species were selected based on data availability for typical body shapes in representative anuran
114
species and life stages.
115
For R. temporaria the geometric mean values of constants from equations for females,
116
males and metamorphs were used. The final values used in the allometric equation were based on
117
the geometric mean of the constants for the three amphibian species.
118
In the studies employing overspray, animals were exposed dorsally while, in the case of
119
filter paper experiments, animals were in contact with impregnated filter paper and were thus
120
exposed ventrally. For the calculations, the exposed skin surface area was taken as either the
121
ventral or the dorsal surface, but not both. The exposed skin area of an anuran, dorsally or ventrally,
6
ACCEPTED MANUSCRIPT
122
can be approximated by an ellipse (see Figure 1) with area A = π ∙ a [major axis] ∙ b [minor axis],
123
in which a = SVL / 2 (with SVL in cm) and b = a / 1.5, and thus:
124
A = π ∙ SVL2 / 6
125
For calculating the internal dose, it was assumed, as a worst case, that 100% of the test substance
126
reaching the skin was absorbed (see ‘Discussion’ section for a detailed examination of this
127
assumption). Finally, the LD50 (mg a.s./kg bw) was calculated based on the LR50 (g a.s./ha), the
128
exposed area A (cm2), and the body weight bw (g):
129
LD50, amphibian = LR50, amphibian ∙ A / (bw ∙ 100)
130 131
2.2 Acute toxicity data for fish
132
Acute toxicity values for fish to match the test items (formulated product or active substance) used
133
in the toxicity studies with amphibians were identified from the literature. Acute toxicity data for
134
fish (96-h LC50) on formulated products were collated from the corresponding Material Safety
135
Data Sheets (MSDSs) or from EU draft assessment reports (DARs) (http://dar.efsa.europa.eu/dar-
136
web/provision).
137
(http://dar.efsa.europa.eu/dar-web/provision) or from the ECOTOX database (U.S. EPA, 2017),
138
both accessed May 12, 2017. Rainbow trout (Oncorhynchus mykiss) data were selected
139
preferentially because this cold-water fish species is generally considered to be among the most
140
sensitive fish species to chemicals (Dyer et al. 1997; EFSA 2005), and for this reason O. mykiss is
141
the only species required for acute testing under the new EU Plant Protection Products Regulation
142
(EC 2009). Furthermore, O. mykiss is probably the most common fish species used in PPPs
143
regulatory testing, worldwide. It is therefore predestined as a surrogate for amphibians when
144
assessing toxicity of PPPs. Tests that reported measured concentrations were selected
Data
on
the
active
substances
7
were
obtained
from
EU DARs
ACCEPTED MANUSCRIPT
145
preferentially. Unbounded values and studies in which the water temperature exceeded 18°C were
146
excluded. If more than one 96-h LC50 value was available for a chemical, the geometric mean of
147
these values was used.
148 149
2.3 Fish bioconcentration data
150
Fish bioconcentration factors (BCFs) for active substances were collected from EU DARs
151
(http://dar.efsa.europa.eu/dar-web/provision),
152
(http://www.safe.nite.go.jp/japan/db.html)
153
(https://cfpub.epa.gov/ecotox). A BCF value was included if the study was conducted according
154
to the following criteria cf. OECD TG 305: aqueous exposure; use of juvenile fish; no mortality;
155
minimum test duration 28 d. Unbounded values were excluded. Most BCF studies employ 14C-
156
labelled test compound and in this case uptake and elimination is generally based on total 14C and
157
the BCF translated into parent equivalents. Ideally, extraction and characterization of 14C residues
158
in both water and fish to determine the BCF as parent is done at steady state; however, this is not
159
always the case, and many studies reported BCF values based on total 14C only. Where possible,
160
the BCF was related to the parent compound measured in both water and fish and, additionally,
161
normalised to a lipid content of 5% as suggested by OECD TG 305. In case multiple BCF values
162
were obtained, the geometric mean was used. If no experimental BCF value was available, the
163
BCF was calculated using the BCFBAF™ model v3.01 in EPI Suite™ (U.S. EPA, 2012).
the and
the
Japanese U.S.
EPA
MITI
database
ECOTOX
database
164 165
2.4 Estimation of LD50 for fish
166
LC50 values for fish were converted to LD50 values by using the lethal body burden (LBB) (aka
167
critical body residue, CBR) approach (McCarty and Mackay 1993; Nendza et al. 1997). In this
8
ACCEPTED MANUSCRIPT
168
approach, the LC50 for fish (mg a.s./L) is multiplied by the BCF (L/kg bw) to estimate an LD50
169
(mg a.s./kg bw). As acute LC50 values for fish are determined after a standard exposure duration
170
of 4 d, the BCF used in the LD50 calculations should also reflect a 4 d exposure. For substances for
171
which 4 d is too short to reach equilibrium, BCF values were corrected to 4 d following the
172
recommendations of Feijtel et al. (1997) and Wen et al. (2015). For several studies, the time to
173
reach steady state was reported (see Table 2). If this was not the case, the time to reach 95% of
174
steady state was estimated according to the equations provided in OECD TG 305 (OECD 2012)
175
based on the logKow (values presented in Table 2). The time to reach 95% of steady state is
176
calculated by t95 = 3 / kd where the depuration constant kd (d-1) is derived from logKow using the
177
following equation:
178
log[kd] = 1.47 - 0.414 ∙ logKow
179
If t95 ≤ 4 d, the BCF does not require correction. However, if t95 is > 4 d, then the 96-h BCF is
180
calculated using a first-order kinetic equation (OECD 2012) with t = 4 d.
181 182
BCFt = BCFss (1 - e-kd ∙ t) Finally, the BCF4 d was used to estimate the LD50, fish (mg a.s./kg bw):
183
LD50, fish = LC50, fish ∙ BCF4 d
184 185
2.5 Statistical analysis
186
The strength of the correlations between LD50, amphibian and LD50, fish values was determined by
187
Spearman’s correlation coefficient (thus making no assumptions on the distribution of the values).
188
Then, LD50 values for fish and amphibians were log-transformed and analysed by means of linear
189
regression. All statistical calculations were performed by GraphPad Prism version 5.04 for
190
Windows, GraphPad Software, San Diego California USA, www.graphpad.com.
9
ACCEPTED MANUSCRIPT
191 192 193
3. RESULTS
194
The literature research yielded 32 acute dermal toxicity experiments for amphibian terrestrial life
195
stages from which LD50 values could be derived. The studies were conducted with 15 different
196
PPPs comprising 13 active substances (four fungicides, two herbicides, and seven insecticides)
197
and one surfactant (see Table 3). The surfactant polyethoxylated tallow amine (POEA) was
198
identified as the formulation ingredient causing the toxicity to amphibians, rather than the
199
herbicidal active substance glyphosate (Moore et al., 2012). The amphibian dermal toxicity studies
200
comprised 15 anuran species from seven families. Additional studies containing amphibian dermal
201
data were identified, but could not be used, due to various reasons, including: too few dose groups;
202
not reporting mortality; no 50% effect value derivable (too high mortality, no mortality);
203
experimental details necessary for interpreting or recalculating toxicity values missing (e.g.
204
Dinehart et al., 2009; Snow and Witmer, 2010; Witmer et al., 2015; Tuttle et al., 2008; Dall and
205
Dawes, 2010).
206
Rainbow trout (O. mykiss) acute toxicity data (LC50, 96 h) for active substances or
207
formulations applied in the amphibian dermal toxicity studies were found in all cases except one:
208
trifloxystrobin (tested as the formulated product Stratego) for which the 96-h LC50 value was
209
derived from a study with bluegill sunfish (Lepomis macrochirus). Experimental BCF values were
210
available for all substances except for POEA, for which the BCF value was calculated. Corrections
211
to the BCF values based on parent compound, lipid content, and to 4 d exposure were made where
212
possible (see Table 2 for details).
10
ACCEPTED MANUSCRIPT
213
Using these data, a highly significant positive correlation between LD50 values for
214
amphibians and fish was found (Spearman rs = 0.943, p < 0.0001). The linear regression of
215
logarithmized LD50 values for amphibians against logarithmized LD50 values for fish resulted in
216
the following interspecies correlation estimation (ICE) model (Raimondo et al., 2007; 2010)
217
equation:
218
log[LD50, amphibian] = 0.852 ∙ log[LC50, fish ∙ BCF4 d] + 0.226
219
The goodness of fit of the regression is characterized by an r² of 0.775 and is shown in Figure 2.
220
It is noted that the regression line suggests that LD50 values for fish and amphibians are quite
221
similar.
222 223
4. DISCUSSION and CONCLUSIONS
224
The ICE model was derived from toxicity data on anurans for chemicals with logKow values
225
ranging from 0.7 (dimethoate) to 6.9 (DDT). It should be noted, that the allometric equation used
226
to relate body length and body weight and the estimation of exposed skin area is only valid for
227
anuran species and not for Urodela that have a different body shape.
228
Considering the overall quality and diversity of the toxicity data for amphibians, the
229
interspecies regression equation derived from log LD50 values for amphibians versus log LD50
230
values for fish gives a reasonably good fit (r² = 0.775). While the (regulatory) LC50 and BCF data
231
for fish are highly standardized, the amphibian data are not, as they are derived from open literature
232
studies using different, often field-collected, test species of varying size and age, employing
233
different exposure designs (i.e. dorsal overspray, ventral contact with filter paper, direct
234
application to the skin), often low numbers of test animals and in many cases, high spacing factors
235
(≥ 10) between treatment levels. Besides the calculated initial body burden received through dorsal 11
ACCEPTED MANUSCRIPT
236
or ventral exposure, amphibians may have absorbed additional substance from subsequent contact
237
with the substrate (e.g. soil or filter paper) in their exposure container; a potential contribution
238
which could not be quantified. In most studies on amphibians, there was no chemical analysis to
239
confirm the correct exposure levels. Furthermore, the assumption that the dermally received dose
240
is completely absorbed probably overestimates body residues. For instance, Shah et al. (1983)
241
found in grass frog (Rana pipiens) body burdens of 85% for parathion, 96% for carbaryl, 41% for
242
DDT, 23% for dieldrin, and 56% for permethrin in relation to the total dose absorbed after a local
243
dorsal application of the test substance. Also for pyraclostrobin and metconazole (the active
244
ingredients in Headline AMP), significantly lower concentrations (5% and 9%, respectively) were
245
measured in frog tissue than expected based on the estimated received dose (Cusaac et al., 2016).
246
The authors hypothesised that this was due to metabolism, but no data (e.g. measured metabolites)
247
to support this were presented. In addition, the estimated exposed skin area used by Cusaac et al.
248
(2017) was larger than calculated in this paper and the frogs were rinsed before analysis. In in vitro
249
studies, 69-83% of the insecticide malathion was absorbed through the skin of American bullfrog
250
(Rana catesbeiana) and the cane toad (Bufo marinus) (Willens et al., 2006) with a tendency for
251
higher absorption through ventral skin than for dorsal skin. All in all, the LD50 values for
252
amphibians are less robust/bear more uncertainty, than the LD50 values for fish.
253
Further, when studying the graph presented in Figure 2, there appear some potential
254
outliers. There are three data points below the 95% confidence bands of the regression line (PPPs
255
containing the actives fenoxaprop-p-ethyl, spiroxamine, and toxaphene) and two data points above
256
the 95% confidence bands of the regression line (for PPPs containing the actives trifloxystrobin,
257
and pyraclostrobin; but only when tested as Headline AMP). For the data below the line, the model
258
would overestimate the LD50 for amphibians and for the ones above the line the LD50 for 12
ACCEPTED MANUSCRIPT
259
amphibians would be understimated. In conclusion, these ‘outliers’ are merely considered to
260
reflect some aspects of variability in the amphibian data, due to (a combination of) factors
261
discussed above. Further, although the fish data are more robust than the amphibian data, they also
262
bear some uncertainties. For instance, not all BCF studies could be recalculated to parent substance
263
or to a fish lipid content of 5% (see Table 2). The BCF for toxaphene was the only value not
264
derived from a regulatory standard test but from the open literature (U.S. EPA ECOTOX database,
265
https://cfpub.epa.gov/ecotox), which may therefore show some deviations compared to the
266
regulatory study values. An interesting case is the organophosphate insecticide malathion for
267
which the BCF in bluegill sunfish based on parent substance could not be quantified, but is
268
significantly
269
http://dar.efsa.europa.eu/dar-web/provision). Also, tiger salamanders exposed to soil spiked with
270
malathion showed a low soil-to-amphibian BCF value of maximally 0.133 (Henson-Ramsey et al.,
271
2008). This BCF value was based on a summation of measured concentrations for malathion and
272
its primary, insecticidally active, metabolite malaoxon. All in all, the available data suggest that
273
malathion BCF values in fish as well as in amphibians are much lower if related to parent
274
compound, but it would need better data to make a more informed decision.
lower
than
the
reported
14C-based
BCF
of
103
L/kg
(EU
DAR,
275
A potential refinement of the estimation method could be the use of acute aquatic
276
amphibian (i.e. tadpole) 96-h LC50 and BCF data, instead of fish data. This would avoid the
277
‘taxonomic jump’ from fish to amphibians and it might improve the predictions. However, aquatic
278
amphibian data (i.e. 96-h LC50 and BCF) are not standardized in contrast to the highly standardized
279
regulatory LC50 values for fish (Weltje et al., 2013) and BCF values for fish, thereby introducing
280
other sources of variation related to differences in test design. In any case, LC50 or BCF data for
281
amphibians are not part of the regulatory requirements and are thus not readily available and 13
ACCEPTED MANUSCRIPT
282
certainly not for new compounds or for the various formulations. In contrast, acute fish data are
283
generated for all formulations (with very few exceptions) as these are required for risk assessment
284
and for classification and labelling purposes. Fish BCF studies with active substance are triggered,
285
typically if logKow > 3, but in case not available, can be calculated by means of QSARs (e.g.
286
BCFBAF™ model in EPI Suite™, U.S. EPA, 2012). A QSAR for prediction of bioconcentration
287
in amphibians is presently not available.
288
For dieldrin, a detailed comparison of LD50 values between fish and aquatic and terrestrial
289
life stages of amphibians is possible, based on the data in Tables 2 and 3 and the toxicity and
290
bioconcentration studies by Schuytema et al. (1991) in which tadpoles of Xenopus laevis (African
291
clawed frog) and Rana catesbeiana (bullfrog) were exposed. The obtained (geometric mean) 96-h
292
LC50 values for X. laevis and R. catesbeiana were 44.72 and 16.24 µg/L, respectively. The 96-h
293
LD50 values, based on measured tissue concentrations for X. laevis and R. catesbeiana were 11.0
294
and 8.6 mg/kg, respectively. (Geometric mean) 4-d BCF values for X. laevis and R. catesbeiana,
295
were 240 and 569 L/kg, respectively, based on measurements in the lowest exposure
296
concentrations of the acute tests. The lowest concentrations were selected, as BCF studies are
297
typically conducted at low non-toxic concentrations. Multiplying the 96-h LC50 with the 4-d BCF
298
to obtain an estimate of the LD50 results in values of 10.56 and 9.24 mg/kg for X. laevis and
299
R. catesbeiana, respectively. Long-term (28-d) bioconcentration tests with tadpoles were also
300
conducted and resulted in steady-state geometric mean BCF values of 927 and 781 L/kg for
301
X. laevis and R. catesbeiana, respectively. Correcting the 28-d BCF values to 4-d BCF values
302
(using the equation described in the Materials and Methods section) would yield of 523 and 441
303
L/kg, respectively. Another difference is the lipid content, which was around 1% in the tadpoles.
14
ACCEPTED MANUSCRIPT
304
The values for dieldrin used in this paper are an LD50 of 12.5 mg/kg for fish, based on a
305
rainbow trout LC50 of 2.0 µg/L, and a 4-d BCF in carp of 5424 L/kg (for primary sources and
306
further details see Table 2). For dermally exposed terrestrial life stages of amphibians, three LD50
307
values of 9.83, 12.97 and 15.16 mg/kg were obtained (for primary sources and further details see
308
Table 3). Apparently, rainbow trout is more sensitive than the tadpoles of the tested amphibian
309
species (following the general trend that in most cases fish are more sensitive than aquatic
310
amphibians as described by Weltje et al. (2013)). Also, fish show higher BCF values, likely
311
because they process a higher volume of water through their gills than tadpoles. However, the
312
measured LD50, but also the product of the LC50 and the 4-d BCF, is very similar between
313
amphibians and fish. The correspondence between the two LD50 values indicates that the LBB
314
values for dieldrin in fish and terrestrial and aquatic amphibians are very similar and supports the
315
developed ICE model.
316
As the interspecies correlation equation is derived from a relatively limited dataset of 32
317
experiments, it would be desirable to update the regression when further amphibian data become
318
available (assuming fish data is always available). Such an update may result in a slight change of
319
the regression values, but these are not expected to change dramatically. What might happen is
320
that substances are identified that are outside the domain of the regression equation; compounds
321
that are acting differently on fish than they are on dermally exposed terrestrial life stages of
322
amphibians. One such compound appears to be the piscicide rotenone for which a field study
323
described that all gill-breathing tadpoles of Columbia spotted frogs (Rana luteiventris) were killed
324
within 24 hours after application, but non-gill-breathing metamorphs, juveniles and adult
325
amphibians were unaffected (Billman et al., 2012). A very rough calculation on terrestrial life
326
stages of bullfrog (Rana catesbeiana), dermally exposed to a rotenone formulation, yielded 40% 15
ACCEPTED MANUSCRIPT
327
mortality at an estimated body burden of ca. 125 mg/kg (Witmer et al., 2015), which is much
328
greater than what would be calculated for fish. Therefore, the outcome of the ICE model based on
329
fish data would have been conservative in the case of rotenone.
330
In this paper, it was shown that by combining existing concepts and approaches from the
331
eco(toxico)logical literature, it was possible to establish a relationship between fish critical body
332
residues and terrestrial amphibian dermal toxicity values. Thus, by making use of standard
333
regulatory fish data (i.e. LC50 and BCF), a prediction of acute toxicity due to dermal exposure of
334
terrestrial life stages of amphibians is possible without additional testing of vertebrates (i.e.
335
amphibians). For risk assessment purposes, the estimated amphibian dermal LD50 is then converted
336
to an LR50 by using the allometric equation. Finally, the obtained LR50 can be compared to a PPP
337
application rate to obtain a risk quotient, enabling a distinction between PPP uses of low concern
338
and those of potential concern. The possible use of this method as a screening approach for risk
339
assessment of PPPs, including considerations on exposure, tiered testing, assessment factors and
340
mitigation measures is detailed in Weltje et al. (submitted). The presented non-testing screening
341
method therefore constitutes a major step forward in advancing the risk assessment for PPPs and
342
terrestrial life stages of amphibians.
343 344 345
Acknowledgements
346
The authors wish to thank Keith Solomon (Guelph University, Canada) for providing additional
347
information on the Glyphos formulation study, Mick Hamer (Syngenta, UK), Alan Samel (DuPont,
348
USA), Gunnar Schmidt, Andreas Ufer (BASF SE, Germany) and four anonymous reviewers for
16
ACCEPTED MANUSCRIPT
349
constructive comments, and Thomas Preuss (Bayer CropScience) and James Wheeler (Dow
350
AgroSciences, UK) for stimulating discussions.
351 352 353
Conflict of interest
354
LW, PJ and PS work for chemical companies as disclosed by their affiliations and some of the
355
substances discussed in this paper are products of these companies.
356 357 358
REFERENCES
359
Aldrich,
360
A.P.,
2009.
Empfindlichkeit
von
Amphibien
gegenüber
Pflanzenschutzmitteln.
AGRARForschung 16, 466–471.
361
Bancila, R.I., Hartel, T., Plaiasu, R., Smets, J., Cogalniceanu, D., 2010. Comparing three body condition
362
indices in amphibians: a case study of yellow-bellied toad Bombina variegate. Amphibia-Reptilia
363
31, 558-562.
364 365
Belden, J., McMurry, S., Smith, L., Reilley, P., 2010. Acute toxicity of fungicide formulations to amphibians at environmentally relevant concentrations. Environ. Toxicol. Chem. 29, 2477-2480.
366
Berger, G., Graef, F., Bethwell, C., Brühl, C.A., Alscher, A., Schmidt, T., Weber, B., 2015. Assessment of
367
pesticide exposure of amphibians and reptiles in agricultural landscapes in Germany and evaluation
368
of the present pesticide risk assessment practice in EU. In: Protection of Biodiversity in the Risk
369
Assessment and Risk Management of Pesticides (Plant Protection Products & Biocides) with a
370
Focus on Arthropods, Soil Organisms and Amphibians. TEXTE 76/2015, p. 82-159,
371
Umweltbundesamt, Dessau-Rosslau, Germany.
17
ACCEPTED MANUSCRIPT
372
Bernal, M.H., Solomon, K.R., Carrasquilla, G., 2009. Toxicity of formulated glyphosate (Glyphos) and
373
cosmo-flux to larval and juvenile Colombian frogs 2. Field and laboratory microcosm acute
374
toxicity. J. Toxicol. Environ. Health, Part A 72, 966-973.
375
Billman, H.G., Kruse, C.G., St-Hilaire, S., Koel, T.M., Arnold, J.L., Peterson, C.R., 2012. Effects of
376
rotenone on Columbia spotted frogs Rana luteiventris during field applications in lentic habitats of
377
Southwestern Montana. N. Am. J. Fish. Manage. 32, 781-789.
378 379 380 381
Brühl, C.A., Pieper, S., Weber, B. 2011., Amphibians at risk? Susceptibility of terrestrial amphibian life stages to pesticides. Environ. Toxicol. Chem. 30, 2465-2472. Brühl, C.A., Schmidt, T., Pieper, S., Alscher, A., 2013. Terrestrial pesticide exposure of amphibians: An underestimated cause of global decline? Sci. Rep. 3, 1135.
382
Crane, M., Finnegan, M., Weltje, L., Kosmala-Grzechnik, S., Gross, M., Wheeler, J.R., 2016. Acute oral
383
toxicity of chemicals in terrestrial life stages of amphibians: Comparisons to birds and mammals.
384
Regul. Toxicol. Pharm. 80, 335-341.
385
Cusaac, J.P.W., Morrison, S.A., Belden, J.B., Smith, L.M., McMurry, S.T., 2016. Acute toxicity of
386
Headline® fungicide to Blanchard’s cricket frogs (Acris blanchardi). Ecotoxicology 25, 447-455.
387
Cusaac, J.P.W., Mimbs, W.H., Belden, J.B., Smith, L.M., McMurry, S.T., 2017. Factors influencing the
388
toxicity of Headline® fungicides to terrestrial stage toads. Environ. Toxicol. Chem.
389
doi:10.1002/etc.3816.
390
Dinehart, S.K., Smith, L.M., McMurry, S.T., Anderson, T.A., Smith, P.N., Haukos, D.A., 2009. Toxicity
391
of a glufosinate- and several glyphosate-based herbicides to juvenile amphibians from the Southern
392
High Plains, USA. Sci. Total Environ. 407, 1065-1071.
393 394
Dyer, S.D., Belanger, S.E., Carr, G.J., 1997. An initial evaluation of the use of Euro/North American fish species for tropical effects assessments. Chemosphere 35, 2767–2781.
395
EC, 2009. Regulation (EC) No. 1107/2009 of the European Parliament and of the Council of 21 October
396
2009 Concerning the Placing of Plant Protection Products on the Market and Repealing Council
397
Directives 79/117/EEC and 91/414/EEC. European Commission, Brussels, Belgium. 18
ACCEPTED MANUSCRIPT
398
EFSA, 2005. Opinion of The Scientific Panel on Plant Health, Plant Protection Products and Their Residues
399
on a Request from EFSA Related to the Assessment of the Acute and Chronic Risk to Aquatic
400
Organisms with Regard to The Possibility of Lowering the Uncertainty Factor If Additional Species
401
Were Tested. The EFSA Journal 301, 1-45
402 403
EFSA 2013. Guidance on Tiered Risk Assessment for Plant Protection Products for Aquatic Organisms in Edge-of-Field Surface Waters. EFSA Journal 11, 3290.
404
Feijtel, T., Kloepper-Sams, P., van den Haan, K., van Egmond, R., Comber, M., Heusel, R. Wierich, P. Ten
405
Berge, W. Gard, A. de Wolf, W., Niessen, H., 1997. Integration of bioaccumulation in an
406
environmental risk assessment. Chemosphere 34, 2337-2350.
407 408
Ferguson, D.E., Gilbert, C.C., 1967. Tolerances of three species of anuran amphibians to five chlorinated hydrocarbon insecticides. J. Miss. Acad. Sci. 13, 135-138.
409
Fryday, S., Thompson, H., 2012. Toxicity of Pesticides to Aquatic and Terrestrial Life Stages of
410
Amphibians and Occurrence, Habitat Use and Exposure of Amphibian Species in Agricultural
411
Environments. EFSA Supporting Publications EN-343 pp. 348.
412 413 414 415
Gibbons, M.M., McCarthy, T.K., 1984. Growth, maturation and survival of frogs Rana temporaria L. Holarctic Ecol. 7, 419-427. Grözinger, F., Thein, J., Feldhaar, H., Rödel, M.-O., 2014. Giants, dwarfs and the environment – metamorphic trait plasticity in the common frog. PLOS ONE 9, e89982.
416
Henson-Ramsey, H., Kennedy-Stoskopf, S., Levine, J., Taylor, S., Shea, D., Stoskopf, M., 2008. Acute
417
toxicity and tissue distributions of malathion in Ambystoma tigrinum. Arch. Environ. Contam.
418
Toxicol. 55, 481-487.
419 420 421 422
Katagi, T., Ose, K., 2014. Bioconcentration and metabolism of pesticides and industrial chemicals in the frog. J. Pest. Sci. 39, 55-68. Kuhn, J., 1994. Lebensgeschichte und Demographie von Erdkrötenweibchen Bufo bufo bufo (L.). Z. Feldherpetologie 1, 3-87.
19
ACCEPTED MANUSCRIPT
423
Llewelyn, V.K., Berger L., Glass B.D., 2016. Percutaneous absorption of chemicals: developing an
424
understanding for the treatment of disease in frogs. J. Vet. Pharmacol. Ther. 39, 109-121.
425
McCarty, L.S., MacKay, D., 1993. Enhancing ecotoxicological modeling and assessment. Body residues
426
and modes of toxic action. Environ. Sci. Technol. 27, 1718–1728.
427
Moore, L.J., Fuentes, L., Rodgers, J.H. Jr, Bowerman, W.W., Yarrow, G.K., Chao, W.Y., Bridges, W.C.
428
Jr., 2012. Relative toxicity of the components of the original formulation of Roundup to five North
429
American anurans. Ecotox. Environ. Safe. 78, 128–133.
430 431 432 433
Nendza, M., Herbst, T., Kussatz, C., Gies A., 1997. Potential for secondary poisoning and biomagnification in marine organisms. Chemosphere 9, 1875-1885. Quaranta, A., Bellantuono V., Cassano G., Lippe C., 2009. Why amphibians are more sensitive than mammals to xenobiotics. PLoS ONE 4, 1-4.
434
Raimondo, S., Jackson, C.R., Barron, M.G., 2010. Influence of taxonomic relatedness and chemical mode
435
of action in acute interspecies estimation models for aquatic species. Environ. Sci. Technol. 44,
436
7711-7716.
437 438 439 440 441 442
Raimondo, S., Mineau, P., Barron, M.G., 2007. Estimation of chemical toxicity to wildlife species using interspecies correlation models. Environ. Sci. Technol. 41, 5888-5894. Shah, P.V., Monroe, R.J., Gathrie, F.E., 1983. Comparative penetration of insecticides in target and nontarget species. Drug Chem. Toxicol. 6, 155–179. Smith, P.N., Cobb, G.P., Godard-Codding, C., Hoff, D., McMurry, S.T., Rainwater, T.R., Reynolds, K.D., 2007. Contaminant exposure in terrestrial vertebrates. Environ. Pollut. 150, 41-64.
443
Snow, N.P., Witmer, G., 2010. American bullfrogs as invasive species: A review of the introduction,
444
subsequent problems, management options, and future directions. Proc. 24th Vertebr. Pest Conf.
445
(R.M. Timm and K.A. Fagerstone, Eds.) Published at Univ. of Calif., Davis. Pp. 86-89.
446 447
Taylor, S.K., Williams, E.S., Mills, K.W., 1999. Effects of malathion on disease susceptibility in Woodhouse’s toads. J. Wildlife Dis. 35, 536–541.
20
ACCEPTED MANUSCRIPT
448 449
Tuttle, N.C., A, Beard. K.H., Al-Chokhachy, R., 2008. Aerially applied citric acid reduces the density of an invasive frog in Hawaii, USA. Wildlife Res. 35, 676-683.
450
U.S. EPA, 2012. Estimation Programs Interface Suite™ for Microsoft® Windows, v 4.11. United States
451
Environmental Protection Agency, Washington, DC, USA. Available: https://www.epa.gov/tsca-
452
screening-tools/epi-suitetm-estimation-program-interface.
453 454
U.S. EPA, 2017. ECOTOX User Guide: ECOTOXicology Knowledgebase System. Version 4.0. Available: http:/www.epa.gov/ecotox/.
455
Dall, D.J., Dawes, J., 2010. U.S. Patent Application US 2010/0069506 A1. Composition for pest control.
456
Appl. No. 12/312,500. Available: http://www.freepatentsonline.com/y2010/0069506.html.
457
Vinson, S.B., Boyd, C.E., Ferguson, D.E., 1963. Aldrin toxicity and cross-resistance in cricket frog.
458
Herpetologica 19, 77-80.
459
Weltje, L., Simpson, P., Gross, M., Crane, M., Wheeler, J.R., 2013. Comparative acute and chronic
460
sensitivity of fish and amphibians: a critical review of data. Environ. Toxicol. Chem. 32, 984-994.
461
Weltje L., Ufer A., Hamer M., Sowig P., Demmig S., Dechet F., (submitted) Considerations on risk
462
assessment for plant protection products for terrestrial life-stages of amphibians.
463
Wen, Y., Su, L., Qin, W., Zhao, Y., Madden, J.C., Steinmetz, F.P., Cronin, M.T.D., 2015. Investigation of
464
critical body residues and modes of toxic action based on injection and aquatic exposure in fish.
465
Water Air Soil Poll. 226, 174.
466
Willens S., Stoskopf, M.K., Baynes, R.E., Lewbart, G.A., Taylor, S.K., Kennedy-Stoskopf, S., 2006.
467
Percutaneous malathion absorption by anuran skin in flow-through diffusion cells. Environ.
468
Toxicol. Phar. 22, 255–262.
469 470
Witmer, G.W., Snow, N.P., Moulton, R.S., 2015. Efficacy of potential chemical control compounds for removing invasive American bullfrogs (Rana catesbeiana). SpringerPlus 4, 497.
471
21
ACCEPTED MANUSCRIPT
472 473
Figure 1. Approximating the exposed skin area of an anuran amphibian by an ellipse (a = major axis; b =
474
minor axis).
475
22
ACCEPTED MANUSCRIPT
log[LD50, amphibian]
2.4 1.8 1.2 0.6 0.0 0.0
0.6
1.2
1.8
2.4
log[LD50, fish] 476 477
Figure 2. Linear regression with 95% confidence bands of log[LD50] in amphibians against log[LD50] in
478
fish. Symbols represent mean and standard error of the mean (SEM).
479
23
ACCEPTED MANUSCRIPT
480
Table 1. Parameter estimates for three anuran species using the allometric relation between snout-vent-
481
length, SVL [mm] and body weight [g]: log (body weight) = a * log (SVL) - b a
b
source
Bombina variegatus
2.89
3.94
Adults and juveniles; Bancila et al. (2010)
Bufo bufo
2.94
3.90
Females; Kuhn (1994)
Rana temporaria
3.09
4.04
geomean of adults and metamorphs
gravid females 3.19
4.36
Gibbons and McCarthy (1984)
spawned females 2.95
4.10
Gibbons and McCarthy (1984)
all females 3.07
4.23
geomean of gravid and spawned females
males 2.44
3.01
Gibbons and McCarthy (1984)
adults 2.74
3.57
geomean of all females and males
metamorphs 3.49
4.58
derived from data of Grözinger et al. (2014)
overall geometric mean
2.97
3.96
482
24
Table 2. Fish BCF, LC50 data and calculated LD50 values. NR: not reported; NQ: not quantifiable O. mykiss BCF5%
BCF Substance
Species (% lipid content) [L/kg]
lipid
BCFparent
steady state
[L/kg]
at 4 d?
logKow
t95 [d]
LD50, fish
BCF4 d
LC50
[mg
[L/kg]h
[mg
a.s./kg
a.s./L]
bw]
Test item
[L/kg]
Aldrin
13748a
Cyprinus carpio (7)
9820
NR
NR
6.50e
49.9
2099
0.004b
TGAI
8.86
DDT
25139a
Cyprinus carpio (5)
25139
NR
NR
6.90e
73.1
3807
0.006b
TGAI
20.9
Dieldrin
13463a
Cyprinus carpio (7)
9616
NR
NR
5.20e
14.5
5424
0.002b
TGAI
12.5
Endrin
10684a
Cyprinus carpio (7)
7632
NR
NR
5.20e
14.5
4305
0.0003b
TGAI
1.42
Toxaphene
25043b
--
NR
NR
5.80e
25.6
9370
0.006b
TGAI
50.2
Fundulus similis (NR), Cyprinodon variegatus (NR) Bromoxynil230c
Lepomis macrochirus (NR)
--
NR
yes
5.90c
--
0.040j
Curol B
9.26
Captan
126c
Lepomis macrochirus (NR)
--
NR
yes
2.57c
--
0.058j
Captan Omya
7.35
Dimethoate
0.8a
Cyprinus carpio (4.3)
0.9
NR
NR
0.70c
0.2
17.60c
Roxion
15.8
Fenoxaprop-p-ethyl
313c
Lepomis macrochirus (8.95)
175
NQ
yes
4.58c
--
0.277j
Dicomil Ultra
48.5
Malathion
103c
Lepomis macrochirus (NR)
--
NQ
yes
2.75c
--
0.180c
TGAI
18.5
octanoate
25
Pyraclostrobin
696c
Lepomis macrochirus (NR)
--
438
yes
3.99c
--
--
0.006c
Headline AMP
0.005j
Headline 250 EC 2.17
2.63
Spiroxamine
79c
Lepomis macrochirus (NR)
--
11
yes
2.89c, f
--
5.750j
Prosper
61.7
Trifloxystrobin
347c
Lepomis macrochirus (NR)
--
NQ
yes
4.50c
--
0.005i, j
Stratego
1.82
POEA
151d
NA
--
151
--
5.89g
--
0.594j
Roundup Ultra
89.6
aMITI
(http://www.safe.nite.go.jp/japan/db.html); bU.S. EPA ECOTOX database (https://cfpub.epa.gov/ecotox); cEU DAR (http://dar.efsa.europa.eu/dar-web/provision);
dBCFBAF™
model v3.01 in EPI Suite™ (U.S. EPA, 2012); eexperimental data from KOWWIN™ model in EPI suite™ (U.S. EPA, 2012); fmean of two isomers; gcalculated
with KOWWIN™ model in EPI Suite™ (U.S. EPA, 2012); hcalculated based on kd; istudy conducted with Lepomis macrochirus; jMaterial Safety Data Sheet (MSDS).
26
ACCEPTED MANUSCRIPT
Body length (mm)
Exposed skin area (cm2)
385
1.06
22.0
2.53
9.17
Aldrin
I
Acris crepitans
Filter paper
1.5 5
446
1.19
22.8
2.73
10.7
Aldrin
I
Filter paper
1.5 5
280
1.13
22.4
2.63
6.55
Filter paper
1.5 5
975
0.93
21.0
2.31
24.3
Filter paper
1.5 5
853
0.92
21.0
2.30
21.2
Species
LD50 (mg a.s./kg bw)
Body weight (g)
1.5 5
Referenceb
Filter paper
Study duration (d)
Acris gryllus
Method of exposure
I
Use typea
Aldrin
Test item
LR50 (g a.s./ha)
Table 3. Amphibian dermal toxicity data and calculated LD50 values
Bufo woodhousei fowleric DDT
I
DDT
I
Acris crepitans Bufo woodhousei fowleric
Dieldrin
I
Acris gryllus
Filter paper
1.5 5
545
1.06
22.0
2.53
13.0
Dieldrin
I
Acris crepitans
Filter paper
1.5 5
660
1.19
22.8
2.73
15.2
Dieldrin
I
Filter paper
1.5 5
398
0.95
21.2
2.36
9.83
Bufo woodhousei fowleric Endrin
I
Acris gryllus
Filter paper
1.5 5
47.2
1.06
22.0
2.53
1.12
Endrin
I
Acris crepitans
Filter paper
1.5 5
75.7
1.19
22.8
2.73
1.74
Endrin
I
Filter paper
1.5 5
85.4
1.06
22.0
2.53
2.03
Filter paper
1.5 5
786
0.93
21.0
2.31
19.6
Filter paper
1.5 5
919
0.99
21.5
2.41
22.4
Overspray
7
476
0.77
19.7
2.04
12.6
Bufo woodhousei fowleric Toxaphene
I
Toxaphene
I
Acris crepitans Bufo woodhousei fowleric
Bromoxynil-octanoate
H
Rana temporaria
27
3
ACCEPTED MANUSCRIPT (as Curol B) Captan F
Rana temporaria
Overspray
7
3
384
0.73
19.4
1.96
10.4
I
Rana temporaria
Overspray
7
3
600
0.74
19.5
1.98
16.1
H
Rana temporaria
Overspray
7
3
528
0.93
21.0
2.31
13.2
I
Bufo woodhouseid
5
6
/
/
/
/
16.5
(as Captan Omya) Dimethoate (as Roxion) Fenoxaprop-p-ethyl (as Dicomil) Ventral Malathion
application Pyraclostrobin F
Acris blanchardi
Overspray
4
4
170
0.49
19.4
1.97
6.78
F
Bufo cognatuse
Overspray
4
7
212
0.85
20.7
2.24
5.60
F
Bufo cognatuse
Overspray
3
1
147
1.06
22.0
2.53
3.49
F
Rana temporaria
Overspray
7
3
65.7
0.93
21.0
2.31
1.64
F
Acris blanchardi
Overspray
4
4
210
0.49
19.4
1.97
8.38
F
Rana temporaria
Overspray
7
3
747
0.89
20.7
2.25
18.8
F
Bufo cognatuse
Overspray
3
1
348
1.06
22.0
2.53
8.29
Overspray
4
2
1398 0.25
13.5
0.96
53.4
(as Headline AMP) Pyraclostrobin (as Headline AMP) Pyraclostrobin (as Headline) Pyraclostrobin (as Headline) Pyraclostrobin (as Headline) Spiroxamine (as Prosper) Trifloxystrobin (as Stratego) POEA (in Glyphos + Cosmo-
Centrolene S
Flux)
prosoblepon
28
ACCEPTED MANUSCRIPT POEA (in Glyphos + Cosmo-
Pristimantis S
Overspray
4
2
1740 0.85
20.4
2.18
44.5
S
Rhinella granulosa Overspray
4
2
2020 0.06
8.5
0.38
121
S
Scinax ruber
Overspray
4
2
2268 0.18
12.1
0.77
96.2
S
Rhinella typhonius
Overspray
4
2
4599 0.05
7.7
0.31
303
Overspray
4
2
6090 0.18
12.0
0.75
262
Overspray
4
2
7085 0.11
10.4
0.56
350
Flux)
taeniatus
POEA (in Glyphos + CosmoFlux) POEA (in Glyphos + CosmoFlux) POEA (in Glyphos + CosmoFlux) POEA (in Glyphos + Cosmo-
Engystomops S
Flux)
pustulosus
POEA (in Glyphos + CosmoS
Rhinella marina
Flux) ause
type: I = insecticide, F = fungicide, H = herbicide, S = surfactant; bReferences: 1 = Belden et al. (2010), 2 =
Bernal et al. (2009), 3 = Brühl et al. (2013), 4 = Cusaac et al. (2016), 5 = Ferguson and Gilbert (1967), 6 = Taylor et al. (1999), 7 = Cusaac et al. (2017); crenamed Anaxyrus fowleri; drenamed Anaxyrus woodhousii; erenamed
Anaxyrus cognatus.
29
Lennart Weltje, Philipp Janz, Peter Sowig PII:
S0045-6535(17)31462-5
DOI:
10.1016/j.chemosphere.2017.09.047
Reference:
CHEM 19918
To appear in:
Chemosphere
Received Date:
23 June 2017
Revised Date:
10 September 2017
Accepted Date:
11 September 2017
Please cite this article as: Lennart Weltje, Philipp Janz, Peter Sowig, An interspecies correlation model to predict acute dermal toxicity of plant protection products to terrestrial life stages of amphibians using fish acute toxicity and bioconcentration data, Chemosphere (2017), doi: 10.1016 /j.chemosphere.2017.09.047
This is a PDF file of an unedited manuscript that has been accepted for publication. As a service to our customers we are providing this early version of the manuscript. The manuscript will undergo copyediting, typesetting, and review of the resulting proof before it is published in its final form. Please note that during the production process errors may be discovered which could affect the content, and all legal disclaimers that apply to the journal pertain.
ACCEPTED MANUSCRIPT Highlights for the manuscript: AN INTERSPECIES CORRELATION MODEL TO PREDICT ACUTE DERMAL TOXICITY OF PLANT PROTECTION PRODUCTS TO TERRESTRIAL LIFE STAGES OF AMPHIBIANS Lennart Weltje, Philipp Janz and Peter Sowig
1. 2. 3. 4.
An equation is derived to predict acute dermal toxicity to terrestrial amphibians Fish bioconcentration and acute toxicity data are required as input The model can be used in a screening approach for pesticide risk assessment Application would significantly reduce amphibian (vertebrate) testing
ACCEPTED MANUSCRIPT
1
AN INTERSPECIES CORRELATION MODEL TO PREDICT ACUTE DERMAL
2
TOXICITY OF PLANT PROTECTION PRODUCTS TO TERRESTRIAL LIFE STAGES
3
OF AMPHIBIANS USING FISH ACUTE TOXICITY AND BIOCONCENTRATION
4
DATA
5 6
Lennart Weltjea*, Philipp Janza, and Peter Sowigb
7 8
a
9
Germany
BASF SE, Crop Protection – Ecotoxicology, Speyerer-Strasse 2, D-67117 Limburgerhof,
10
b
11
* author for correspondence. E-mail: [email protected], Tel.: +49 – 62160 - 28564
Bayer CropScience AG, Industriepark Höchst, D-65926 Frankfurt-Höchst, Germany
1
ACCEPTED MANUSCRIPT
12
ABSTRACT
13
This paper presents a model to predict acute dermal toxicity of plant protection products (PPPs) to
14
terrestrial amphibian life stages from (regulatory) fish data. By combining existing concepts,
15
including interspecies correlation estimation (ICE), allometric relations, lethal body burden (LBB)
16
and bioconcentration modelling, an equation was derived that predicts the amphibian median lethal
17
dermal dose (LD50) from standard acute toxicity values (96-h LC50) for fish and bioconcentration
18
factors (BCF) in fish. Where possible, fish BCF values were corrected to 5% lipid, and to parent
19
compound. Then, BCF values were adjusted to an exposure duration of 96 hours, in case steady
20
state took longer to be achieved. The derived correlation equation is based on 32 LD50 values from
21
acute dermal toxicity experiments with 15 different species of anuran amphibians, comprising 15
22
different PPPs. The developed ICE model can be used in a screening approach to estimate the
23
acute risk to amphibian terrestrial life stages from dermal exposures to PPPs with organic active
24
substances. This has the potential to reduce unnecessary testing of vertebrates.
25 26 27 28
Key words: terrestrial toxicity, amphibians, dermal exposure, fish toxicity, risk assessment, lethal
29
body burden; alternative method
2
ACCEPTED MANUSCRIPT
30
1. INTRODUCTION
31
The European Regulation 1107/2009 concerning the placing of plant protection products (PPPs)
32
on the market (EC, 2009) requires the risks for amphibians to be assessed based on available
33
information. Similar wording can be found in the European Food and Safety Authority (EFSA)
34
Guidance Document for risk assessment for aquatic organisms in edge-of-the-field waters (EFSA,
35
2013). Currently, amphibians are not explicitly assessed as it is assumed that the risk assessment
36
for other vertebrates covers the risk to amphibians (both aquatic and terrestrial life stages).
37
However, concerns have recently been voiced regarding potential risks of PPPs to amphibian
38
terrestrial life stages that might not be covered by the current risk assessment (e.g. Belden et al.,
39
2010; Brühl et al., 2011). Against this background, the EFSA has been given a mandate to develop
40
a guidance document for the risk assessment of PPPs for amphibians (and reptiles [note that this
41
paper focuses on amphibians only]). This is a challenging task, given that there are no standard
42
methods available for describing and quantifying amphibian exposure and toxicity. In addition,
43
the life-cycle of amphibians requires consideration of both the aquatic and terrestrial life stages.
44
The situation is further complicated by the demand to minimize testing of vertebrates due to animal
45
welfare (EC Regulation 1107/2009). Therefore, as a first approach it makes sense to consider if
46
the current data requirements on aquatic and terrestrial vertebrates yield data that can be used as
47
surrogates to inform on the sensitivity of amphibians to PPPs.
48
While comprehensive data reviews have shown that the acute and chronic sensitivities of
49
aquatic life stages of amphibians are similar to or lower than those of fish (Aldrich, 2009; Weltje
50
et al., 2013) and thus can be covered in the risk assessment, the situation is less clear for the
51
terrestrial life stages. In general, there are fewer data available on terrestrial than on aquatic life
52
stages. Nevertheless, a recent data comparison of oral acute toxicity values for 26 chemicals
3
ACCEPTED MANUSCRIPT
53
showed that the sensitivity of amphibian terrestrial life stages can be covered by available standard
54
toxicity data on birds and mammals (Crane et al., 2016). However, the exposure route that is
55
probably of highest relevance for terrestrial life stages of amphibians is the dermal exposure route
56
(Fryday and Thompson, 2012; Smith et al., 2007). In contrast to other terrestrial vertebrates,
57
amphibians do not have fur and feathers that might serve as a protective barrier against chemical
58
exposure (Smith et al., 2007). Furthermore, amphibians have a more permeable skin (Quaranta et
59
al., 2009) fulfilling physiological functions such as water and gas exchange (i.e. breathing). In
60
situations where exposure to a PPPs could occur, for instance during a spraying application in a
61
cropped field, this permeability may lead to dermal uptake that is typically not experienced by
62
other vertebrate taxa (Smith et al., 2007). Because of the different properties of the skin,
63
mammalian and avian species are probably not suitable surrogates for dermal exposure routes in
64
the amphibian risk assessment. Accordingly, Berger et al. (2015) found no statistically significant
65
correlation between the acute toxicity of PPPs to amphibians after dermal exposure and the
66
potential of PPPs for skin irritation in rabbits. Therefore, the work presented here aimed at
67
developing a method for predicting acute dermal toxicity in terrestrial amphibian life stages, based
68
on other available data. Such a non-testing method is urgently needed for use in a screening
69
approach to distinguish PPPs of lower from those of higher potential concern. Only when a
70
potential concern is identified, a PPP requires further consideration, which may eventually involve
71
in vivo vertebrate toxicity testing using amphibians. A non-testing approach would thus improve
72
current procedures for risk assessment and avoid unnecessary testing of vertebrates.
73
To derive an equation for acute dermal toxicity estimation in terrestrial life stages of
74
amphibians, some assumptions were made and various concepts established in (aquatic)
75
ecotoxicology were combined. In short, the high correlation between sensitivity of fish to
4
ACCEPTED MANUSCRIPT
76
chemicals and that of amphibian aquatic life stages (Weltje et al., 2013) was assumed to translate
77
to the terrestrial life stage. Further, the lethal body burden (LBB) concept was used to convert LC50
78
values for fish, by multiplication with the bioconcentration factor (BCF), to LD50 values. For
79
dermally exposed amphibians, the LD50 values were calculated by considering the exposed skin
80
surface, body weight, and the exposure rate. The dermally received dose was assumed to be
81
absorbed completely by the exposed amphibians. More detailed explanations are provided in the
82
‘Materials and Methods’ section.
83 84 85
2. MATERIALS AND METHODS
86
2.1 Acute dermal toxicity data for amphibians
87
Acute dermal toxicity data for amphibian terrestrial life stages were collected from Fryday and
88
Thompson (2012) and targeted searches of the scientific literature via Web of Science
89
(http://webofknowledge.com/WOS accessed May 12, 2017) using (amphibian* OR frog* OR
90
toad* OR anura* OR urodel* OR caudat*) AND (dermal OR terrestrial OR overspray). Dermal
91
toxicity data for terrestrial life stages of amphibians are generated mainly for three purposes: i) to
92
study side-effects of chemicals in ecotoxicology (e.g. Vinson et al., 1963); ii) to study chemical
93
means to control invasive amphibian species for animal conservation purposes (e.g. Witmer et al.,
94
2015); iii) to study side-effects of veterinary medicines if applied percutaneously (an alternative
95
to oral application, e.g. Llewelyn et al., 2015). Only studies with dermal exposure routes (i.e.
96
overspray, filter paper, direct application) employing at least 2 dose groups and reporting mortality
97
as an endpoint were included. When formulated PPPs were tested, the application rates were
98
converted to active substance (a.s.). Taylor et al. (1999), Belden et al. (2010) and Brühl et al.
5
ACCEPTED MANUSCRIPT
99
(2013) did not report 50% lethal exposure levels. To obtain these values, they were calculated for
100
each of the tested chemicals from the original data in consideration of: i) the observed mortality at
101
each treatment, ii) any observed control mortality, and iii) the high spacing factor of 10 used in
102
these studies. If more than one 50% lethal effect value for a species was available (e.g. Ferguson
103
and Gilbert, 1967), the geometric mean of these values was used.
104
For converting dermal exposure values to body burdens, the exposed body surface and the
105
body weight of the experimental animals are needed. Therefore, allometric equations were applied
106
for relating exposed skin area to body length and body weight in anurans. If body weight (bw) or
107
body length (i.e. snout to vent length (SVL)) were not reported, SVL (in mm) was estimated based
108
on bw (in g) or vice versa, using the following allometric equation:
109
log[bw] = 2.97 ∙ log[SVL] - 3.96
110
The numerical parameter values were derived from allometric equations for yellow-bellied toad
111
Bombina variegata (Bancila et al. 2010), common toad Bufo bufo (Kuhn 1994), and common frog
112
Rana temporaria (Gibbons and McCarthy 1984; Grözinger et al. 2014) – see Table 1. These
113
species were selected based on data availability for typical body shapes in representative anuran
114
species and life stages.
115
For R. temporaria the geometric mean values of constants from equations for females,
116
males and metamorphs were used. The final values used in the allometric equation were based on
117
the geometric mean of the constants for the three amphibian species.
118
In the studies employing overspray, animals were exposed dorsally while, in the case of
119
filter paper experiments, animals were in contact with impregnated filter paper and were thus
120
exposed ventrally. For the calculations, the exposed skin surface area was taken as either the
121
ventral or the dorsal surface, but not both. The exposed skin area of an anuran, dorsally or ventrally,
6
ACCEPTED MANUSCRIPT
122
can be approximated by an ellipse (see Figure 1) with area A = π ∙ a [major axis] ∙ b [minor axis],
123
in which a = SVL / 2 (with SVL in cm) and b = a / 1.5, and thus:
124
A = π ∙ SVL2 / 6
125
For calculating the internal dose, it was assumed, as a worst case, that 100% of the test substance
126
reaching the skin was absorbed (see ‘Discussion’ section for a detailed examination of this
127
assumption). Finally, the LD50 (mg a.s./kg bw) was calculated based on the LR50 (g a.s./ha), the
128
exposed area A (cm2), and the body weight bw (g):
129
LD50, amphibian = LR50, amphibian ∙ A / (bw ∙ 100)
130 131
2.2 Acute toxicity data for fish
132
Acute toxicity values for fish to match the test items (formulated product or active substance) used
133
in the toxicity studies with amphibians were identified from the literature. Acute toxicity data for
134
fish (96-h LC50) on formulated products were collated from the corresponding Material Safety
135
Data Sheets (MSDSs) or from EU draft assessment reports (DARs) (http://dar.efsa.europa.eu/dar-
136
web/provision).
137
(http://dar.efsa.europa.eu/dar-web/provision) or from the ECOTOX database (U.S. EPA, 2017),
138
both accessed May 12, 2017. Rainbow trout (Oncorhynchus mykiss) data were selected
139
preferentially because this cold-water fish species is generally considered to be among the most
140
sensitive fish species to chemicals (Dyer et al. 1997; EFSA 2005), and for this reason O. mykiss is
141
the only species required for acute testing under the new EU Plant Protection Products Regulation
142
(EC 2009). Furthermore, O. mykiss is probably the most common fish species used in PPPs
143
regulatory testing, worldwide. It is therefore predestined as a surrogate for amphibians when
144
assessing toxicity of PPPs. Tests that reported measured concentrations were selected
Data
on
the
active
substances
7
were
obtained
from
EU DARs
ACCEPTED MANUSCRIPT
145
preferentially. Unbounded values and studies in which the water temperature exceeded 18°C were
146
excluded. If more than one 96-h LC50 value was available for a chemical, the geometric mean of
147
these values was used.
148 149
2.3 Fish bioconcentration data
150
Fish bioconcentration factors (BCFs) for active substances were collected from EU DARs
151
(http://dar.efsa.europa.eu/dar-web/provision),
152
(http://www.safe.nite.go.jp/japan/db.html)
153
(https://cfpub.epa.gov/ecotox). A BCF value was included if the study was conducted according
154
to the following criteria cf. OECD TG 305: aqueous exposure; use of juvenile fish; no mortality;
155
minimum test duration 28 d. Unbounded values were excluded. Most BCF studies employ 14C-
156
labelled test compound and in this case uptake and elimination is generally based on total 14C and
157
the BCF translated into parent equivalents. Ideally, extraction and characterization of 14C residues
158
in both water and fish to determine the BCF as parent is done at steady state; however, this is not
159
always the case, and many studies reported BCF values based on total 14C only. Where possible,
160
the BCF was related to the parent compound measured in both water and fish and, additionally,
161
normalised to a lipid content of 5% as suggested by OECD TG 305. In case multiple BCF values
162
were obtained, the geometric mean was used. If no experimental BCF value was available, the
163
BCF was calculated using the BCFBAF™ model v3.01 in EPI Suite™ (U.S. EPA, 2012).
the and
the
Japanese U.S.
EPA
MITI
database
ECOTOX
database
164 165
2.4 Estimation of LD50 for fish
166
LC50 values for fish were converted to LD50 values by using the lethal body burden (LBB) (aka
167
critical body residue, CBR) approach (McCarty and Mackay 1993; Nendza et al. 1997). In this
8
ACCEPTED MANUSCRIPT
168
approach, the LC50 for fish (mg a.s./L) is multiplied by the BCF (L/kg bw) to estimate an LD50
169
(mg a.s./kg bw). As acute LC50 values for fish are determined after a standard exposure duration
170
of 4 d, the BCF used in the LD50 calculations should also reflect a 4 d exposure. For substances for
171
which 4 d is too short to reach equilibrium, BCF values were corrected to 4 d following the
172
recommendations of Feijtel et al. (1997) and Wen et al. (2015). For several studies, the time to
173
reach steady state was reported (see Table 2). If this was not the case, the time to reach 95% of
174
steady state was estimated according to the equations provided in OECD TG 305 (OECD 2012)
175
based on the logKow (values presented in Table 2). The time to reach 95% of steady state is
176
calculated by t95 = 3 / kd where the depuration constant kd (d-1) is derived from logKow using the
177
following equation:
178
log[kd] = 1.47 - 0.414 ∙ logKow
179
If t95 ≤ 4 d, the BCF does not require correction. However, if t95 is > 4 d, then the 96-h BCF is
180
calculated using a first-order kinetic equation (OECD 2012) with t = 4 d.
181 182
BCFt = BCFss (1 - e-kd ∙ t) Finally, the BCF4 d was used to estimate the LD50, fish (mg a.s./kg bw):
183
LD50, fish = LC50, fish ∙ BCF4 d
184 185
2.5 Statistical analysis
186
The strength of the correlations between LD50, amphibian and LD50, fish values was determined by
187
Spearman’s correlation coefficient (thus making no assumptions on the distribution of the values).
188
Then, LD50 values for fish and amphibians were log-transformed and analysed by means of linear
189
regression. All statistical calculations were performed by GraphPad Prism version 5.04 for
190
Windows, GraphPad Software, San Diego California USA, www.graphpad.com.
9
ACCEPTED MANUSCRIPT
191 192 193
3. RESULTS
194
The literature research yielded 32 acute dermal toxicity experiments for amphibian terrestrial life
195
stages from which LD50 values could be derived. The studies were conducted with 15 different
196
PPPs comprising 13 active substances (four fungicides, two herbicides, and seven insecticides)
197
and one surfactant (see Table 3). The surfactant polyethoxylated tallow amine (POEA) was
198
identified as the formulation ingredient causing the toxicity to amphibians, rather than the
199
herbicidal active substance glyphosate (Moore et al., 2012). The amphibian dermal toxicity studies
200
comprised 15 anuran species from seven families. Additional studies containing amphibian dermal
201
data were identified, but could not be used, due to various reasons, including: too few dose groups;
202
not reporting mortality; no 50% effect value derivable (too high mortality, no mortality);
203
experimental details necessary for interpreting or recalculating toxicity values missing (e.g.
204
Dinehart et al., 2009; Snow and Witmer, 2010; Witmer et al., 2015; Tuttle et al., 2008; Dall and
205
Dawes, 2010).
206
Rainbow trout (O. mykiss) acute toxicity data (LC50, 96 h) for active substances or
207
formulations applied in the amphibian dermal toxicity studies were found in all cases except one:
208
trifloxystrobin (tested as the formulated product Stratego) for which the 96-h LC50 value was
209
derived from a study with bluegill sunfish (Lepomis macrochirus). Experimental BCF values were
210
available for all substances except for POEA, for which the BCF value was calculated. Corrections
211
to the BCF values based on parent compound, lipid content, and to 4 d exposure were made where
212
possible (see Table 2 for details).
10
ACCEPTED MANUSCRIPT
213
Using these data, a highly significant positive correlation between LD50 values for
214
amphibians and fish was found (Spearman rs = 0.943, p < 0.0001). The linear regression of
215
logarithmized LD50 values for amphibians against logarithmized LD50 values for fish resulted in
216
the following interspecies correlation estimation (ICE) model (Raimondo et al., 2007; 2010)
217
equation:
218
log[LD50, amphibian] = 0.852 ∙ log[LC50, fish ∙ BCF4 d] + 0.226
219
The goodness of fit of the regression is characterized by an r² of 0.775 and is shown in Figure 2.
220
It is noted that the regression line suggests that LD50 values for fish and amphibians are quite
221
similar.
222 223
4. DISCUSSION and CONCLUSIONS
224
The ICE model was derived from toxicity data on anurans for chemicals with logKow values
225
ranging from 0.7 (dimethoate) to 6.9 (DDT). It should be noted, that the allometric equation used
226
to relate body length and body weight and the estimation of exposed skin area is only valid for
227
anuran species and not for Urodela that have a different body shape.
228
Considering the overall quality and diversity of the toxicity data for amphibians, the
229
interspecies regression equation derived from log LD50 values for amphibians versus log LD50
230
values for fish gives a reasonably good fit (r² = 0.775). While the (regulatory) LC50 and BCF data
231
for fish are highly standardized, the amphibian data are not, as they are derived from open literature
232
studies using different, often field-collected, test species of varying size and age, employing
233
different exposure designs (i.e. dorsal overspray, ventral contact with filter paper, direct
234
application to the skin), often low numbers of test animals and in many cases, high spacing factors
235
(≥ 10) between treatment levels. Besides the calculated initial body burden received through dorsal 11
ACCEPTED MANUSCRIPT
236
or ventral exposure, amphibians may have absorbed additional substance from subsequent contact
237
with the substrate (e.g. soil or filter paper) in their exposure container; a potential contribution
238
which could not be quantified. In most studies on amphibians, there was no chemical analysis to
239
confirm the correct exposure levels. Furthermore, the assumption that the dermally received dose
240
is completely absorbed probably overestimates body residues. For instance, Shah et al. (1983)
241
found in grass frog (Rana pipiens) body burdens of 85% for parathion, 96% for carbaryl, 41% for
242
DDT, 23% for dieldrin, and 56% for permethrin in relation to the total dose absorbed after a local
243
dorsal application of the test substance. Also for pyraclostrobin and metconazole (the active
244
ingredients in Headline AMP), significantly lower concentrations (5% and 9%, respectively) were
245
measured in frog tissue than expected based on the estimated received dose (Cusaac et al., 2016).
246
The authors hypothesised that this was due to metabolism, but no data (e.g. measured metabolites)
247
to support this were presented. In addition, the estimated exposed skin area used by Cusaac et al.
248
(2017) was larger than calculated in this paper and the frogs were rinsed before analysis. In in vitro
249
studies, 69-83% of the insecticide malathion was absorbed through the skin of American bullfrog
250
(Rana catesbeiana) and the cane toad (Bufo marinus) (Willens et al., 2006) with a tendency for
251
higher absorption through ventral skin than for dorsal skin. All in all, the LD50 values for
252
amphibians are less robust/bear more uncertainty, than the LD50 values for fish.
253
Further, when studying the graph presented in Figure 2, there appear some potential
254
outliers. There are three data points below the 95% confidence bands of the regression line (PPPs
255
containing the actives fenoxaprop-p-ethyl, spiroxamine, and toxaphene) and two data points above
256
the 95% confidence bands of the regression line (for PPPs containing the actives trifloxystrobin,
257
and pyraclostrobin; but only when tested as Headline AMP). For the data below the line, the model
258
would overestimate the LD50 for amphibians and for the ones above the line the LD50 for 12
ACCEPTED MANUSCRIPT
259
amphibians would be understimated. In conclusion, these ‘outliers’ are merely considered to
260
reflect some aspects of variability in the amphibian data, due to (a combination of) factors
261
discussed above. Further, although the fish data are more robust than the amphibian data, they also
262
bear some uncertainties. For instance, not all BCF studies could be recalculated to parent substance
263
or to a fish lipid content of 5% (see Table 2). The BCF for toxaphene was the only value not
264
derived from a regulatory standard test but from the open literature (U.S. EPA ECOTOX database,
265
https://cfpub.epa.gov/ecotox), which may therefore show some deviations compared to the
266
regulatory study values. An interesting case is the organophosphate insecticide malathion for
267
which the BCF in bluegill sunfish based on parent substance could not be quantified, but is
268
significantly
269
http://dar.efsa.europa.eu/dar-web/provision). Also, tiger salamanders exposed to soil spiked with
270
malathion showed a low soil-to-amphibian BCF value of maximally 0.133 (Henson-Ramsey et al.,
271
2008). This BCF value was based on a summation of measured concentrations for malathion and
272
its primary, insecticidally active, metabolite malaoxon. All in all, the available data suggest that
273
malathion BCF values in fish as well as in amphibians are much lower if related to parent
274
compound, but it would need better data to make a more informed decision.
lower
than
the
reported
14C-based
BCF
of
103
L/kg
(EU
DAR,
275
A potential refinement of the estimation method could be the use of acute aquatic
276
amphibian (i.e. tadpole) 96-h LC50 and BCF data, instead of fish data. This would avoid the
277
‘taxonomic jump’ from fish to amphibians and it might improve the predictions. However, aquatic
278
amphibian data (i.e. 96-h LC50 and BCF) are not standardized in contrast to the highly standardized
279
regulatory LC50 values for fish (Weltje et al., 2013) and BCF values for fish, thereby introducing
280
other sources of variation related to differences in test design. In any case, LC50 or BCF data for
281
amphibians are not part of the regulatory requirements and are thus not readily available and 13
ACCEPTED MANUSCRIPT
282
certainly not for new compounds or for the various formulations. In contrast, acute fish data are
283
generated for all formulations (with very few exceptions) as these are required for risk assessment
284
and for classification and labelling purposes. Fish BCF studies with active substance are triggered,
285
typically if logKow > 3, but in case not available, can be calculated by means of QSARs (e.g.
286
BCFBAF™ model in EPI Suite™, U.S. EPA, 2012). A QSAR for prediction of bioconcentration
287
in amphibians is presently not available.
288
For dieldrin, a detailed comparison of LD50 values between fish and aquatic and terrestrial
289
life stages of amphibians is possible, based on the data in Tables 2 and 3 and the toxicity and
290
bioconcentration studies by Schuytema et al. (1991) in which tadpoles of Xenopus laevis (African
291
clawed frog) and Rana catesbeiana (bullfrog) were exposed. The obtained (geometric mean) 96-h
292
LC50 values for X. laevis and R. catesbeiana were 44.72 and 16.24 µg/L, respectively. The 96-h
293
LD50 values, based on measured tissue concentrations for X. laevis and R. catesbeiana were 11.0
294
and 8.6 mg/kg, respectively. (Geometric mean) 4-d BCF values for X. laevis and R. catesbeiana,
295
were 240 and 569 L/kg, respectively, based on measurements in the lowest exposure
296
concentrations of the acute tests. The lowest concentrations were selected, as BCF studies are
297
typically conducted at low non-toxic concentrations. Multiplying the 96-h LC50 with the 4-d BCF
298
to obtain an estimate of the LD50 results in values of 10.56 and 9.24 mg/kg for X. laevis and
299
R. catesbeiana, respectively. Long-term (28-d) bioconcentration tests with tadpoles were also
300
conducted and resulted in steady-state geometric mean BCF values of 927 and 781 L/kg for
301
X. laevis and R. catesbeiana, respectively. Correcting the 28-d BCF values to 4-d BCF values
302
(using the equation described in the Materials and Methods section) would yield of 523 and 441
303
L/kg, respectively. Another difference is the lipid content, which was around 1% in the tadpoles.
14
ACCEPTED MANUSCRIPT
304
The values for dieldrin used in this paper are an LD50 of 12.5 mg/kg for fish, based on a
305
rainbow trout LC50 of 2.0 µg/L, and a 4-d BCF in carp of 5424 L/kg (for primary sources and
306
further details see Table 2). For dermally exposed terrestrial life stages of amphibians, three LD50
307
values of 9.83, 12.97 and 15.16 mg/kg were obtained (for primary sources and further details see
308
Table 3). Apparently, rainbow trout is more sensitive than the tadpoles of the tested amphibian
309
species (following the general trend that in most cases fish are more sensitive than aquatic
310
amphibians as described by Weltje et al. (2013)). Also, fish show higher BCF values, likely
311
because they process a higher volume of water through their gills than tadpoles. However, the
312
measured LD50, but also the product of the LC50 and the 4-d BCF, is very similar between
313
amphibians and fish. The correspondence between the two LD50 values indicates that the LBB
314
values for dieldrin in fish and terrestrial and aquatic amphibians are very similar and supports the
315
developed ICE model.
316
As the interspecies correlation equation is derived from a relatively limited dataset of 32
317
experiments, it would be desirable to update the regression when further amphibian data become
318
available (assuming fish data is always available). Such an update may result in a slight change of
319
the regression values, but these are not expected to change dramatically. What might happen is
320
that substances are identified that are outside the domain of the regression equation; compounds
321
that are acting differently on fish than they are on dermally exposed terrestrial life stages of
322
amphibians. One such compound appears to be the piscicide rotenone for which a field study
323
described that all gill-breathing tadpoles of Columbia spotted frogs (Rana luteiventris) were killed
324
within 24 hours after application, but non-gill-breathing metamorphs, juveniles and adult
325
amphibians were unaffected (Billman et al., 2012). A very rough calculation on terrestrial life
326
stages of bullfrog (Rana catesbeiana), dermally exposed to a rotenone formulation, yielded 40% 15
ACCEPTED MANUSCRIPT
327
mortality at an estimated body burden of ca. 125 mg/kg (Witmer et al., 2015), which is much
328
greater than what would be calculated for fish. Therefore, the outcome of the ICE model based on
329
fish data would have been conservative in the case of rotenone.
330
In this paper, it was shown that by combining existing concepts and approaches from the
331
eco(toxico)logical literature, it was possible to establish a relationship between fish critical body
332
residues and terrestrial amphibian dermal toxicity values. Thus, by making use of standard
333
regulatory fish data (i.e. LC50 and BCF), a prediction of acute toxicity due to dermal exposure of
334
terrestrial life stages of amphibians is possible without additional testing of vertebrates (i.e.
335
amphibians). For risk assessment purposes, the estimated amphibian dermal LD50 is then converted
336
to an LR50 by using the allometric equation. Finally, the obtained LR50 can be compared to a PPP
337
application rate to obtain a risk quotient, enabling a distinction between PPP uses of low concern
338
and those of potential concern. The possible use of this method as a screening approach for risk
339
assessment of PPPs, including considerations on exposure, tiered testing, assessment factors and
340
mitigation measures is detailed in Weltje et al. (submitted). The presented non-testing screening
341
method therefore constitutes a major step forward in advancing the risk assessment for PPPs and
342
terrestrial life stages of amphibians.
343 344 345
Acknowledgements
346
The authors wish to thank Keith Solomon (Guelph University, Canada) for providing additional
347
information on the Glyphos formulation study, Mick Hamer (Syngenta, UK), Alan Samel (DuPont,
348
USA), Gunnar Schmidt, Andreas Ufer (BASF SE, Germany) and four anonymous reviewers for
16
ACCEPTED MANUSCRIPT
349
constructive comments, and Thomas Preuss (Bayer CropScience) and James Wheeler (Dow
350
AgroSciences, UK) for stimulating discussions.
351 352 353
Conflict of interest
354
LW, PJ and PS work for chemical companies as disclosed by their affiliations and some of the
355
substances discussed in this paper are products of these companies.
356 357 358
REFERENCES
359
Aldrich,
360
A.P.,
2009.
Empfindlichkeit
von
Amphibien
gegenüber
Pflanzenschutzmitteln.
AGRARForschung 16, 466–471.
361
Bancila, R.I., Hartel, T., Plaiasu, R., Smets, J., Cogalniceanu, D., 2010. Comparing three body condition
362
indices in amphibians: a case study of yellow-bellied toad Bombina variegate. Amphibia-Reptilia
363
31, 558-562.
364 365
Belden, J., McMurry, S., Smith, L., Reilley, P., 2010. Acute toxicity of fungicide formulations to amphibians at environmentally relevant concentrations. Environ. Toxicol. Chem. 29, 2477-2480.
366
Berger, G., Graef, F., Bethwell, C., Brühl, C.A., Alscher, A., Schmidt, T., Weber, B., 2015. Assessment of
367
pesticide exposure of amphibians and reptiles in agricultural landscapes in Germany and evaluation
368
of the present pesticide risk assessment practice in EU. In: Protection of Biodiversity in the Risk
369
Assessment and Risk Management of Pesticides (Plant Protection Products & Biocides) with a
370
Focus on Arthropods, Soil Organisms and Amphibians. TEXTE 76/2015, p. 82-159,
371
Umweltbundesamt, Dessau-Rosslau, Germany.
17
ACCEPTED MANUSCRIPT
372
Bernal, M.H., Solomon, K.R., Carrasquilla, G., 2009. Toxicity of formulated glyphosate (Glyphos) and
373
cosmo-flux to larval and juvenile Colombian frogs 2. Field and laboratory microcosm acute
374
toxicity. J. Toxicol. Environ. Health, Part A 72, 966-973.
375
Billman, H.G., Kruse, C.G., St-Hilaire, S., Koel, T.M., Arnold, J.L., Peterson, C.R., 2012. Effects of
376
rotenone on Columbia spotted frogs Rana luteiventris during field applications in lentic habitats of
377
Southwestern Montana. N. Am. J. Fish. Manage. 32, 781-789.
378 379 380 381
Brühl, C.A., Pieper, S., Weber, B. 2011., Amphibians at risk? Susceptibility of terrestrial amphibian life stages to pesticides. Environ. Toxicol. Chem. 30, 2465-2472. Brühl, C.A., Schmidt, T., Pieper, S., Alscher, A., 2013. Terrestrial pesticide exposure of amphibians: An underestimated cause of global decline? Sci. Rep. 3, 1135.
382
Crane, M., Finnegan, M., Weltje, L., Kosmala-Grzechnik, S., Gross, M., Wheeler, J.R., 2016. Acute oral
383
toxicity of chemicals in terrestrial life stages of amphibians: Comparisons to birds and mammals.
384
Regul. Toxicol. Pharm. 80, 335-341.
385
Cusaac, J.P.W., Morrison, S.A., Belden, J.B., Smith, L.M., McMurry, S.T., 2016. Acute toxicity of
386
Headline® fungicide to Blanchard’s cricket frogs (Acris blanchardi). Ecotoxicology 25, 447-455.
387
Cusaac, J.P.W., Mimbs, W.H., Belden, J.B., Smith, L.M., McMurry, S.T., 2017. Factors influencing the
388
toxicity of Headline® fungicides to terrestrial stage toads. Environ. Toxicol. Chem.
389
doi:10.1002/etc.3816.
390
Dinehart, S.K., Smith, L.M., McMurry, S.T., Anderson, T.A., Smith, P.N., Haukos, D.A., 2009. Toxicity
391
of a glufosinate- and several glyphosate-based herbicides to juvenile amphibians from the Southern
392
High Plains, USA. Sci. Total Environ. 407, 1065-1071.
393 394
Dyer, S.D., Belanger, S.E., Carr, G.J., 1997. An initial evaluation of the use of Euro/North American fish species for tropical effects assessments. Chemosphere 35, 2767–2781.
395
EC, 2009. Regulation (EC) No. 1107/2009 of the European Parliament and of the Council of 21 October
396
2009 Concerning the Placing of Plant Protection Products on the Market and Repealing Council
397
Directives 79/117/EEC and 91/414/EEC. European Commission, Brussels, Belgium. 18
ACCEPTED MANUSCRIPT
398
EFSA, 2005. Opinion of The Scientific Panel on Plant Health, Plant Protection Products and Their Residues
399
on a Request from EFSA Related to the Assessment of the Acute and Chronic Risk to Aquatic
400
Organisms with Regard to The Possibility of Lowering the Uncertainty Factor If Additional Species
401
Were Tested. The EFSA Journal 301, 1-45
402 403
EFSA 2013. Guidance on Tiered Risk Assessment for Plant Protection Products for Aquatic Organisms in Edge-of-Field Surface Waters. EFSA Journal 11, 3290.
404
Feijtel, T., Kloepper-Sams, P., van den Haan, K., van Egmond, R., Comber, M., Heusel, R. Wierich, P. Ten
405
Berge, W. Gard, A. de Wolf, W., Niessen, H., 1997. Integration of bioaccumulation in an
406
environmental risk assessment. Chemosphere 34, 2337-2350.
407 408
Ferguson, D.E., Gilbert, C.C., 1967. Tolerances of three species of anuran amphibians to five chlorinated hydrocarbon insecticides. J. Miss. Acad. Sci. 13, 135-138.
409
Fryday, S., Thompson, H., 2012. Toxicity of Pesticides to Aquatic and Terrestrial Life Stages of
410
Amphibians and Occurrence, Habitat Use and Exposure of Amphibian Species in Agricultural
411
Environments. EFSA Supporting Publications EN-343 pp. 348.
412 413 414 415
Gibbons, M.M., McCarthy, T.K., 1984. Growth, maturation and survival of frogs Rana temporaria L. Holarctic Ecol. 7, 419-427. Grözinger, F., Thein, J., Feldhaar, H., Rödel, M.-O., 2014. Giants, dwarfs and the environment – metamorphic trait plasticity in the common frog. PLOS ONE 9, e89982.
416
Henson-Ramsey, H., Kennedy-Stoskopf, S., Levine, J., Taylor, S., Shea, D., Stoskopf, M., 2008. Acute
417
toxicity and tissue distributions of malathion in Ambystoma tigrinum. Arch. Environ. Contam.
418
Toxicol. 55, 481-487.
419 420 421 422
Katagi, T., Ose, K., 2014. Bioconcentration and metabolism of pesticides and industrial chemicals in the frog. J. Pest. Sci. 39, 55-68. Kuhn, J., 1994. Lebensgeschichte und Demographie von Erdkrötenweibchen Bufo bufo bufo (L.). Z. Feldherpetologie 1, 3-87.
19
ACCEPTED MANUSCRIPT
423
Llewelyn, V.K., Berger L., Glass B.D., 2016. Percutaneous absorption of chemicals: developing an
424
understanding for the treatment of disease in frogs. J. Vet. Pharmacol. Ther. 39, 109-121.
425
McCarty, L.S., MacKay, D., 1993. Enhancing ecotoxicological modeling and assessment. Body residues
426
and modes of toxic action. Environ. Sci. Technol. 27, 1718–1728.
427
Moore, L.J., Fuentes, L., Rodgers, J.H. Jr, Bowerman, W.W., Yarrow, G.K., Chao, W.Y., Bridges, W.C.
428
Jr., 2012. Relative toxicity of the components of the original formulation of Roundup to five North
429
American anurans. Ecotox. Environ. Safe. 78, 128–133.
430 431 432 433
Nendza, M., Herbst, T., Kussatz, C., Gies A., 1997. Potential for secondary poisoning and biomagnification in marine organisms. Chemosphere 9, 1875-1885. Quaranta, A., Bellantuono V., Cassano G., Lippe C., 2009. Why amphibians are more sensitive than mammals to xenobiotics. PLoS ONE 4, 1-4.
434
Raimondo, S., Jackson, C.R., Barron, M.G., 2010. Influence of taxonomic relatedness and chemical mode
435
of action in acute interspecies estimation models for aquatic species. Environ. Sci. Technol. 44,
436
7711-7716.
437 438 439 440 441 442
Raimondo, S., Mineau, P., Barron, M.G., 2007. Estimation of chemical toxicity to wildlife species using interspecies correlation models. Environ. Sci. Technol. 41, 5888-5894. Shah, P.V., Monroe, R.J., Gathrie, F.E., 1983. Comparative penetration of insecticides in target and nontarget species. Drug Chem. Toxicol. 6, 155–179. Smith, P.N., Cobb, G.P., Godard-Codding, C., Hoff, D., McMurry, S.T., Rainwater, T.R., Reynolds, K.D., 2007. Contaminant exposure in terrestrial vertebrates. Environ. Pollut. 150, 41-64.
443
Snow, N.P., Witmer, G., 2010. American bullfrogs as invasive species: A review of the introduction,
444
subsequent problems, management options, and future directions. Proc. 24th Vertebr. Pest Conf.
445
(R.M. Timm and K.A. Fagerstone, Eds.) Published at Univ. of Calif., Davis. Pp. 86-89.
446 447
Taylor, S.K., Williams, E.S., Mills, K.W., 1999. Effects of malathion on disease susceptibility in Woodhouse’s toads. J. Wildlife Dis. 35, 536–541.
20
ACCEPTED MANUSCRIPT
448 449
Tuttle, N.C., A, Beard. K.H., Al-Chokhachy, R., 2008. Aerially applied citric acid reduces the density of an invasive frog in Hawaii, USA. Wildlife Res. 35, 676-683.
450
U.S. EPA, 2012. Estimation Programs Interface Suite™ for Microsoft® Windows, v 4.11. United States
451
Environmental Protection Agency, Washington, DC, USA. Available: https://www.epa.gov/tsca-
452
screening-tools/epi-suitetm-estimation-program-interface.
453 454
U.S. EPA, 2017. ECOTOX User Guide: ECOTOXicology Knowledgebase System. Version 4.0. Available: http:/www.epa.gov/ecotox/.
455
Dall, D.J., Dawes, J., 2010. U.S. Patent Application US 2010/0069506 A1. Composition for pest control.
456
Appl. No. 12/312,500. Available: http://www.freepatentsonline.com/y2010/0069506.html.
457
Vinson, S.B., Boyd, C.E., Ferguson, D.E., 1963. Aldrin toxicity and cross-resistance in cricket frog.
458
Herpetologica 19, 77-80.
459
Weltje, L., Simpson, P., Gross, M., Crane, M., Wheeler, J.R., 2013. Comparative acute and chronic
460
sensitivity of fish and amphibians: a critical review of data. Environ. Toxicol. Chem. 32, 984-994.
461
Weltje L., Ufer A., Hamer M., Sowig P., Demmig S., Dechet F., (submitted) Considerations on risk
462
assessment for plant protection products for terrestrial life-stages of amphibians.
463
Wen, Y., Su, L., Qin, W., Zhao, Y., Madden, J.C., Steinmetz, F.P., Cronin, M.T.D., 2015. Investigation of
464
critical body residues and modes of toxic action based on injection and aquatic exposure in fish.
465
Water Air Soil Poll. 226, 174.
466
Willens S., Stoskopf, M.K., Baynes, R.E., Lewbart, G.A., Taylor, S.K., Kennedy-Stoskopf, S., 2006.
467
Percutaneous malathion absorption by anuran skin in flow-through diffusion cells. Environ.
468
Toxicol. Phar. 22, 255–262.
469 470
Witmer, G.W., Snow, N.P., Moulton, R.S., 2015. Efficacy of potential chemical control compounds for removing invasive American bullfrogs (Rana catesbeiana). SpringerPlus 4, 497.
471
21
ACCEPTED MANUSCRIPT
472 473
Figure 1. Approximating the exposed skin area of an anuran amphibian by an ellipse (a = major axis; b =
474
minor axis).
475
22
ACCEPTED MANUSCRIPT
log[LD50, amphibian]
2.4 1.8 1.2 0.6 0.0 0.0
0.6
1.2
1.8
2.4
log[LD50, fish] 476 477
Figure 2. Linear regression with 95% confidence bands of log[LD50] in amphibians against log[LD50] in
478
fish. Symbols represent mean and standard error of the mean (SEM).
479
23
ACCEPTED MANUSCRIPT
480
Table 1. Parameter estimates for three anuran species using the allometric relation between snout-vent-
481
length, SVL [mm] and body weight [g]: log (body weight) = a * log (SVL) - b a
b
source
Bombina variegatus
2.89
3.94
Adults and juveniles; Bancila et al. (2010)
Bufo bufo
2.94
3.90
Females; Kuhn (1994)
Rana temporaria
3.09
4.04
geomean of adults and metamorphs
gravid females 3.19
4.36
Gibbons and McCarthy (1984)
spawned females 2.95
4.10
Gibbons and McCarthy (1984)
all females 3.07
4.23
geomean of gravid and spawned females
males 2.44
3.01
Gibbons and McCarthy (1984)
adults 2.74
3.57
geomean of all females and males
metamorphs 3.49
4.58
derived from data of Grözinger et al. (2014)
overall geometric mean
2.97
3.96
482
24
Table 2. Fish BCF, LC50 data and calculated LD50 values. NR: not reported; NQ: not quantifiable O. mykiss BCF5%
BCF Substance
Species (% lipid content) [L/kg]
lipid
BCFparent
steady state
[L/kg]
at 4 d?
logKow
t95 [d]
LD50, fish
BCF4 d
LC50
[mg
[L/kg]h
[mg
a.s./kg
a.s./L]
bw]
Test item
[L/kg]
Aldrin
13748a
Cyprinus carpio (7)
9820
NR
NR
6.50e
49.9
2099
0.004b
TGAI
8.86
DDT
25139a
Cyprinus carpio (5)
25139
NR
NR
6.90e
73.1
3807
0.006b
TGAI
20.9
Dieldrin
13463a
Cyprinus carpio (7)
9616
NR
NR
5.20e
14.5
5424
0.002b
TGAI
12.5
Endrin
10684a
Cyprinus carpio (7)
7632
NR
NR
5.20e
14.5
4305
0.0003b
TGAI
1.42
Toxaphene
25043b
--
NR
NR
5.80e
25.6
9370
0.006b
TGAI
50.2
Fundulus similis (NR), Cyprinodon variegatus (NR) Bromoxynil230c
Lepomis macrochirus (NR)
--
NR
yes
5.90c
--
0.040j
Curol B
9.26
Captan
126c
Lepomis macrochirus (NR)
--
NR
yes
2.57c
--
0.058j
Captan Omya
7.35
Dimethoate
0.8a
Cyprinus carpio (4.3)
0.9
NR
NR
0.70c
0.2
17.60c
Roxion
15.8
Fenoxaprop-p-ethyl
313c
Lepomis macrochirus (8.95)
175
NQ
yes
4.58c
--
0.277j
Dicomil Ultra
48.5
Malathion
103c
Lepomis macrochirus (NR)
--
NQ
yes
2.75c
--
0.180c
TGAI
18.5
octanoate
25
Pyraclostrobin
696c
Lepomis macrochirus (NR)
--
438
yes
3.99c
--
--
0.006c
Headline AMP
0.005j
Headline 250 EC 2.17
2.63
Spiroxamine
79c
Lepomis macrochirus (NR)
--
11
yes
2.89c, f
--
5.750j
Prosper
61.7
Trifloxystrobin
347c
Lepomis macrochirus (NR)
--
NQ
yes
4.50c
--
0.005i, j
Stratego
1.82
POEA
151d
NA
--
151
--
5.89g
--
0.594j
Roundup Ultra
89.6
aMITI
(http://www.safe.nite.go.jp/japan/db.html); bU.S. EPA ECOTOX database (https://cfpub.epa.gov/ecotox); cEU DAR (http://dar.efsa.europa.eu/dar-web/provision);
dBCFBAF™
model v3.01 in EPI Suite™ (U.S. EPA, 2012); eexperimental data from KOWWIN™ model in EPI suite™ (U.S. EPA, 2012); fmean of two isomers; gcalculated
with KOWWIN™ model in EPI Suite™ (U.S. EPA, 2012); hcalculated based on kd; istudy conducted with Lepomis macrochirus; jMaterial Safety Data Sheet (MSDS).
26
ACCEPTED MANUSCRIPT
Body length (mm)
Exposed skin area (cm2)
385
1.06
22.0
2.53
9.17
Aldrin
I
Acris crepitans
Filter paper
1.5 5
446
1.19
22.8
2.73
10.7
Aldrin
I
Filter paper
1.5 5
280
1.13
22.4
2.63
6.55
Filter paper
1.5 5
975
0.93
21.0
2.31
24.3
Filter paper
1.5 5
853
0.92
21.0
2.30
21.2
Species
LD50 (mg a.s./kg bw)
Body weight (g)
1.5 5
Referenceb
Filter paper
Study duration (d)
Acris gryllus
Method of exposure
I
Use typea
Aldrin
Test item
LR50 (g a.s./ha)
Table 3. Amphibian dermal toxicity data and calculated LD50 values
Bufo woodhousei fowleric DDT
I
DDT
I
Acris crepitans Bufo woodhousei fowleric
Dieldrin
I
Acris gryllus
Filter paper
1.5 5
545
1.06
22.0
2.53
13.0
Dieldrin
I
Acris crepitans
Filter paper
1.5 5
660
1.19
22.8
2.73
15.2
Dieldrin
I
Filter paper
1.5 5
398
0.95
21.2
2.36
9.83
Bufo woodhousei fowleric Endrin
I
Acris gryllus
Filter paper
1.5 5
47.2
1.06
22.0
2.53
1.12
Endrin
I
Acris crepitans
Filter paper
1.5 5
75.7
1.19
22.8
2.73
1.74
Endrin
I
Filter paper
1.5 5
85.4
1.06
22.0
2.53
2.03
Filter paper
1.5 5
786
0.93
21.0
2.31
19.6
Filter paper
1.5 5
919
0.99
21.5
2.41
22.4
Overspray
7
476
0.77
19.7
2.04
12.6
Bufo woodhousei fowleric Toxaphene
I
Toxaphene
I
Acris crepitans Bufo woodhousei fowleric
Bromoxynil-octanoate
H
Rana temporaria
27
3
ACCEPTED MANUSCRIPT (as Curol B) Captan F
Rana temporaria
Overspray
7
3
384
0.73
19.4
1.96
10.4
I
Rana temporaria
Overspray
7
3
600
0.74
19.5
1.98
16.1
H
Rana temporaria
Overspray
7
3
528
0.93
21.0
2.31
13.2
I
Bufo woodhouseid
5
6
/
/
/
/
16.5
(as Captan Omya) Dimethoate (as Roxion) Fenoxaprop-p-ethyl (as Dicomil) Ventral Malathion
application Pyraclostrobin F
Acris blanchardi
Overspray
4
4
170
0.49
19.4
1.97
6.78
F
Bufo cognatuse
Overspray
4
7
212
0.85
20.7
2.24
5.60
F
Bufo cognatuse
Overspray
3
1
147
1.06
22.0
2.53
3.49
F
Rana temporaria
Overspray
7
3
65.7
0.93
21.0
2.31
1.64
F
Acris blanchardi
Overspray
4
4
210
0.49
19.4
1.97
8.38
F
Rana temporaria
Overspray
7
3
747
0.89
20.7
2.25
18.8
F
Bufo cognatuse
Overspray
3
1
348
1.06
22.0
2.53
8.29
Overspray
4
2
1398 0.25
13.5
0.96
53.4
(as Headline AMP) Pyraclostrobin (as Headline AMP) Pyraclostrobin (as Headline) Pyraclostrobin (as Headline) Pyraclostrobin (as Headline) Spiroxamine (as Prosper) Trifloxystrobin (as Stratego) POEA (in Glyphos + Cosmo-
Centrolene S
Flux)
prosoblepon
28
ACCEPTED MANUSCRIPT POEA (in Glyphos + Cosmo-
Pristimantis S
Overspray
4
2
1740 0.85
20.4
2.18
44.5
S
Rhinella granulosa Overspray
4
2
2020 0.06
8.5
0.38
121
S
Scinax ruber
Overspray
4
2
2268 0.18
12.1
0.77
96.2
S
Rhinella typhonius
Overspray
4
2
4599 0.05
7.7
0.31
303
Overspray
4
2
6090 0.18
12.0
0.75
262
Overspray
4
2
7085 0.11
10.4
0.56
350
Flux)
taeniatus
POEA (in Glyphos + CosmoFlux) POEA (in Glyphos + CosmoFlux) POEA (in Glyphos + CosmoFlux) POEA (in Glyphos + Cosmo-
Engystomops S
Flux)
pustulosus
POEA (in Glyphos + CosmoS
Rhinella marina
Flux) ause
type: I = insecticide, F = fungicide, H = herbicide, S = surfactant; bReferences: 1 = Belden et al. (2010), 2 =
Bernal et al. (2009), 3 = Brühl et al. (2013), 4 = Cusaac et al. (2016), 5 = Ferguson and Gilbert (1967), 6 = Taylor et al. (1999), 7 = Cusaac et al. (2017); crenamed Anaxyrus fowleri; drenamed Anaxyrus woodhousii; erenamed
Anaxyrus cognatus.
29