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Case Study Part 1: How to Calculate Appropriate - ResearchGate

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Ecotoxicology, 14, 877–893, 2005
Г“ 2005 Springer Science+Business Media, Inc. Printed in The U.S.A.
DOI: 10.1007/s10646-005-0034-4
Case Study Part 1: How to Calculate Appropriate Deterministic Long-Term
Toxicity to Exposure Ratios (TERs) for Birds and Mammals
R.F. SHORE,1,* D.R. CROCKER,2 H.R. AKCAKAYA,3 R.S. BENNETT,4 P.F. CHAPMAN,5 M. CLOOK,6
M. CRANE,7 I.C. DEWHURST,7 P.J. EDWARDS,5 A. FAIRBROTHER,8 S. FERSON,3 D. FISCHER,9
A.D.M. HART,2 M. HOLMES,2 M.J. HOOPER,10 M. LAVINE,11 A. LEOPOLD,12 R. LUTTIK,13 P. MINEAU,14
D.R.J. MOORE,15 S.R. MORTENSON,16 D.G. NOBLE,17 R.J. O’CONNOR,18 W. ROELOFS,2 R.M. SIBLY,19
G. C. SMITH,2 M. SPENDIFF,20 T. A. SPRINGER,21 H.M. THOMPSON2 AND C. TOPPING22
1
Centre for Ecology and Hydrology, Monks Wood, Cambridgeshire, UK
2
Central Science Laboratory, Sand Hutton, York, UK
3
Applied Biomathematics, Setauket, NY, USA
4
USEPA/NHEERL/MED, Duluth, MN, USA
5
Jealotts Hill International Research Station, Bracknell, UK
6
Pesticides Safety Directorate, York, UK
7
Crane Consultants, Faringdon, Oxfordshire, UK
8
USEPA/NHEERL/WED, Corvallis, OR, USA
9
Bayer Corp, Research & Development, Stilwell, KS, USA
10
Texas Tech University, Lubbock, TX, USA
11
Duke University, ISDS, Durham, NC, USA
12
Wildlife International, Berkampweg 1, 7231, Warnveld, CL, The Netherlands
13
RIVM, CSR, Utrecht, The Netherlands
14
Canadian Wildlife Service, National Wildlife Research Centre, Ottawa, Canada
15
Ecological Risk Assessment Group, Cadmus Group, Ottawa, Canada
16
Syngenta Crop Protection, Inc., Greensboro, NC, USA
17
British Trust for Ornithology, Norfolk, Thetford, UK
18
Wildlife Ecology, University of Maine, Orono, ME, USA
19
University of Reading, Whiteknights, Reading, UK
20
Health and Safety Laboratory, Broad Lane, Sheffield, UK
21
Wildlife International, Easton, MD, USA
22
EcoSol, FaЛљrupvej 54, DK-8410, RГёnde, Denmark
Accepted 25 March 2005/Published online 23 November 2005
Abstract. In the European Union, п¬Ѓrst-tier assessment of the long-term risk to birds and mammals from
pesticides is based on calculation of a deterministic long-term toxicity/exposure ratio (TERlt). The ratio is
developed from generic herbivores and insectivores and applied to all species. This paper describes two case
studies that implement proposed improvements to the way long-term risk is assessed. These refined
methods require calculation of a TER for each of five identified phases of reproduction (phase-specific
*To whom correspondence should be addressed:
Tel.: +44-1487-772517; Fax: +44-1487-773467;
E-mail: rfs@ceh.ac.uk
878 Shore et al.
TERs) and use of adjusted No Observed Effect Levels (NOELs) to incorporate variation in species sensitivity to pesticides. They also involve progressive refinement of the exposure estimate so that it applies to
particular species, rather than generic indicators, and relates spraying date to onset of reproduction. The
effect of using these new methods on the assessment of risk is described. Each refinement did not necessarily
alter the calculated TER value in a way that was either predictable or consistent across both case studies.
However, use of adjusted NOELs always reduced TERs, and relating spraying date to onset of reproduction increased most phase-specific TERs. The case studies suggested that the current first-tier TERlt
assessment may underestimate risk in some circumstances and that phase-specific assessments can help
identify appropriate risk-reduction measures. The way in which deterministic phase-specific assessments
can currently be implemented to enhance п¬Ѓrst-tier assessment is outlined.
Keywords: risk assessment; pesticide exposure; no observed effect level; skylark; wood mouse
Introduction
The European Union (EU) п¬Ѓrst-tier assessment of
long-term risk to birds and mammals from pesticides (European Commission, 2002) is based on
calculation of a long-term toxicity/exposure ratio
(TERlt); the lower the ratio, the greater the risk to
wildlife. If the TERlt is less than a trigger value of
п¬Ѓve, this indicates that there should be no authorisation of the pesticide unless an appropriate
risk assessment is carried out. The toxicity component of this ratio is derived from the lowest No
Observed Effect Level (NOEL) determined for
birds using the avian reproduction test (OECD,
1998) and for mammals using the suite of endpoints measured in medium and long-term toxicity
tests (European Commission, 2002). The exposure
term is an ��estimated theoretical exposure’’ (ETE)
based on standard crop and wildlife scenarios
using default values for generic indicator species. It
is meant to represent a realistic worst-case assessment.
TERs are a relatively crude measure of risk and
their use has been criticised (Tiebout and Brugger,
1995; Calow, 1998; Bennett et al., 2005; Mineau,
2005). For example, there is disagreement whether
the exposure estimates and the toxicity endpoints
are over or under protective, there is no indication
of the type or scale of effects, there is uncertainty
about the validity and accuracy of extrapolating
laboratory toxicity data to wild species in the
natural environment, and there is no objective
scientific justification for a TERlt trigger value of
п¬Ѓve. Furthermore, the TERlt, in providing a single
deterministic value, tends to mask the variability
and uncertainty that exists in both the exposure
and toxicity data (Solomon et al., 2000).
A workshop was convened in January 2004 with
the aim of improving long-term risk assessment for
birds and mammals (Hart and Thompson, 2005).
Better methods for estimating exposure were outlined (Crocker, 2005) and a novel approach for
assessing toxicity was proposed; appropriate
NOELs were determined for key stages in the
breeding cycle and the possible consequence of
toxic impact at each stage were estimated (Bennett
et al., 2005). Ways of incorporating inter-species
differences in toxicological response into the
assessment were considered (Luttik et al. 2005).
Methods were also developed to link the timing of
reproduction to the time-course of exposure
(Fischer, 2005), and to translate the assessment of
risk to individuals into likely effects on populations
(Sibley et al., 2005). The current paper is п¬Ѓrst of
two that describe case studies which implement the
approaches developed at the workshop. There were
two case studies, one for birds and another for
mammals. Both used specific, but different, exposure scenarios. The aim of this first paper is to
illustrate the way in which deterministic TERs can
be generated for different reproductive phases
(phase-specific TERs), as advocated by Bennett
et al. (2005), and then be progressively refined using
more realistic exposure estimates. The way in
which these approaches can be incorporated into
current risk assessment procedures is also
described. The second case study paper (Roelofs
et al., 2005) outlines a more radical departure from
current deterministic TER assessments. It shows
how probabilistic approaches can be applied to the
Long-Term Toxicity to Exposure Ratios (TERs) for Birds and Mammals 879
production of phase-specific TERs, and how these
can then be developed to model overall impacts on
individual breeding success and on subsequent
population effects.
Description of case study scenarios
Crop type
Six standard scenarios (grassland, early cereals,
late cereals, early/late leafy crops, orchard/vine/
hops, seed treatment) are used for EU risk
assessments (European Commission, 2002). Winter wheat, a cereal crop widely grown in Britain
and that currently occupies 42% of arable land
(Garthwaite et al., 2003), was the scenario that
was taken for both the bird and mammal case
studies.
Indicator species
When using the EU standard scenarios for risk
assessment, the calculated ETE for the п¬Ѓrst-tier
risk assessment is calculated for generic indicators.
In cereals, these are a large herbivorous bird
(goose, unspecified species), a small insectivorous
bird (wren, Troglodytes spp or tit, Parus spp), a
small herbivorous mammal (vole, unspecified
species) and an insectivorous mammal (shrew Sorex spp) (European Commission, 2002). In Britain,
the skylark (Alauda arvensis), an omnivorous bird
that feeds its chicks on insects (Green, 1978;
Poulsen et al., 1998; Donald et al., 2001) and the
wood mouse (Apodemus sylvaticus), a ubiquitous
rodent that eats mostly seeds and insects in arable
systems (Tattersall and Macdonald, 2003), occur
widely in arable п¬Ѓelds where they forage extensively and also breed. Both species are relatively
small (and so have high food intake to body mass
ratios), and are likely to ingest pesticide-contaminated forage. They are therefore likely to be good
specific (rather than generic) indicators for pesticide-related effects on birds and mammals in
cereals systems, and so were used in the case
studies. An added advantage was that the biology
of these species is well studied and the ecological
data required for the case studies was readily
available.
Pesticides in case studies
Although the pesticides in the case study scenarios
were п¬Ѓctitious and deliberately given qualities that
would create long-term risk, all of their toxicological and agronomic attributes were realistic. An
insecticide, compound I, was used in the skylark
case study and was borrowed from a previous case
study (Mineau et al., 2001). Compound I was assumed to be applied once only between April and
July at a rate of 0.25 kg/ha. A fungicide, compound F, was used for the wood mouse case study.
It was assumed that this was applied between
March and July at a rate of 0.75 kg/ha and that
there were two applications separated by seven
days. The timing of spraying for these compounds
reflected typical usage patterns for insecticides and
fungicides in Britain (Garthwaite et al., 2003). The
relevant toxicity data needed for calculation of the
TERs for compounds I and F, are given in
Tables 1 and 2 respectively, together with values
adjusted to take into account inter-species variability in sensitivity (Luttik et al., 2005).
Calculation of toxicity/exposure ratios
Refinement stages of the risk assessment
The case studies begin by using the current EU
guidance to assess the risk of insecticide I to birds
and fungicide F to mammals. They then gradually
increase the realism of the toxicity and exposure
scenarios. The progression is as follows:
1. Standard п¬Ѓrst-tier assessment using generic
wildlife species and default values for pesticide residues and half-life as described in current EU guidance (European Commission,
2002).
2. Phase-specific long-term TERs for default
indicator species. The toxicity endpoints identified by Bennett et al. (2005) for different
reproductive phases are used to calculate
phase–specific TERs. The exposure estimate
is refined so that it is weighted over a time
period appropriate to the endpoint.
3. Phase-specific long-term TERs for skylark
and wood mouse. Generic species are replaced
with wildlife species that are vulnerable to
880 Shore et al.
Table 1. Avian toxicity data for compound I. Data are from Mineau et al. (2001)
a
Toxicity end-point
Acute NOEL for adult body weight prelaying
NOEL for eggs laid per hen per day
NOEL for mean eggshell thickness per hena
NOEL for eggshell cracking
NOEL for % fertile eggs per egg set per hen
NOEL for proportion of hatching per egg set per hen
NOEL for proportion of 14-day chicks per no/hatchlings per hen
NOEL for 14-day chick body weight
d
Five-day dietary LC50
Estimated п¬Ѓve-day LC05
d,e
Test
speciesb
Toxicity
value
Adjusted
toxicity valuec
Unit for
toxicity value
BQ
JQ
M
JQ
M
JQ
M
JQ
M
JQ
M
JQ
M
JQ
M
M
M
20
2.10
0.157
6.66
0.261
6.66
0.261
6.66
0.261
6.66
0.261
2.10
0.157
6.66
0.261
22.1
9.53
0.608
0.029
mg/kg
mg/kg BW/d
0.067
mg/kg BW/d
0.067
mg/kg BW/d
0.067
mg/kg BW/d
0.067
mg/kg BW/d
0.029
mg/kg BW/d
0.067
mg/kg BW/d
0.672
0.289
mg/kg BW/d
mg/kg BW/d
a
No observed effect level (NOELs) for dietary intake (mg/kg BW/d) calculated from reported no observed effect concentrations
(NOECs) using data on the feeding rates and body weights of test animals.
b
Species are BQ: bobwhite quail (Colinus virginianus); JQ: Japanese quail (Coturnix coturnix japonica), M: mallard (Anas platyrynchos).
c
Toxicity factors adjusted to incorporate inter-species extrapolation using factors calculated following Luttik et al. (2005). These values
are used in Tables 4–7.
d
Dietary lethal concentration (LC) data converted to daily intake per unit body weight using data on the feeding rates and body
weights of test animals.
e
Assuming a probit slope of approximately 4.5 (Mastrota, pers. comm.), LC50 may converted to LC05 by multiplying the LC50 value by
0.431.
Table 2. Toxicity data for compound F in rats
NOEL value for various toxicity end-points (mg/kg BW/24 h)a
Toxicity value
Adjusted toxicity valueb
90-day toxicity test
Behaviour in 2-generation testc
no/pups per mated female
teratogenicity
No/weanlings per mated female
Juvenile survival to 4 weeks
Body weight at 2 weeks old
No/weanlings per F2 female
22.4
4.5c
90
30
90
90
27
4.5
1.50
0.302
6.04
2.01
6.04
6.04
1.81
0.302
a
No observed effect level (NOELs) for dietary intake (mg/kg BW/d) either as reported or calculated from no observed effect concentrations (NOECs) using data on the feeding rates and body weights of test animals.
b
Toxicity factors adjusted to incorporate inter-species extrapolation using factors calculated following Luttik et al. (2005). These values
are used in Tables 4–7.
c
Lowest value taken from measurements made on effects on food or water intake or body weight.
exposure due to the proposed use of the pesticide. Exposure calculations make use of
specific body weight, diet composition and
habitat use.
4. Chronological TERs. The species breeding
calendar and spraying dates are applied to
the phase-specific long-term TERs for the
skylark and wood mouse.
Long-Term Toxicity to Exposure Ratios (TERs) for Birds and Mammals 881
Standard п¬Ѓrst-tier assessment of long term toxicity/
exposure ratio (TERlt)
Phase-specific long-term TERs for default
indicators
The п¬Ѓrst-tier TERlt, is calculated as the NOEL/
ETE ratio. The toxicity term used in the calculations was the most sensitive of the range of possible endpoints determined for compound I in
birds (avian case study, Table 1) and compound F
in laboratory mammals (mammalian case study,
Table 2). The NOEL is often calculated from a
dietary No Observed Effect Concentration
(NOEC) using the reported feeding rate and body
weight of the test animals. The ETE is for generic
insectivorous and herbivorous birds and mammals
that have п¬Ѓxed default body weights and food intake rates (Table 3). Calculation of the ETE is
described in detail elsewhere (European Commission, 2002; Crocker, 2005). Default values
(Table 3) are used for all exposure inputs other
than application rate and the number of repeat
applications of the pesticide. It is assumed that
contaminated diet is not avoided (avoidance factor
(AV) = 1), animals feed wholly on a single food
type (proportion of food type in the diet
(PD) = 1) and all diet is from the treated area
(proportion of the diet obtained in the treated area
(PT) = 1).
In the case study scenarios, the п¬Ѓrst-tier TERlt
values for birds and mammals (Table 3) were less
than п¬Ѓve, the trigger value.
Although the п¬Ѓrst-tier TERlt values in the case
studies were all less then п¬Ѓve, they give no indication of the exact nature, likely scale or frequency
of toxic effects. They do not even indicate that one
or more reproductive stage is likely to be affected,
nor indicate a likely adverse effect on the specific
toxicity endpoint used in calculating the TERlt
values. This is, in part, because the default time
period of three weeks (post-application), over
which a time weighted average (TWA) exposure is
calculated, may bear little relation to the exposure
period that is actually critical to the specific toxicity endpoint. It could be argued that any potential underestimate of exposure due to the use of
a TWA exposure factor is more than compensated
for by use of the lowest appropriate NOEL in the
TERlt. The result is likely to be a conservative, and
therefore precautionary, п¬Ѓrst-tier assessment.
Whether this is the case is unproven. If true, it is
probable that acceptable products will be inappropriately rejected.
Bennett et al. (2005) argued that a more objective
and informative assessment of likely effects on
reproduction can be gained by calculating specific
TERs for each phase of reproduction. This also has
the advantage of explicitly aligning the exposure
term to the specific toxicity endpoint. Phase-specific
Table 3. Default values (European Commission, 2002) for the various parameters used to calculate the п¬Ѓrst-tier estimated theoretical exposure (ETE), application rate for pesticides, and the calculated values for the п¬Ѓrst-tier ETE and TERlt for insecticide I and
birds and for fungicide F and mammals in a cereals scenario
Indicator species
Compound I
Insectivorous bird (10 g)
Large herbivorous bird (3000 g)
Compound F
Insectivorous mammal (10 g)
Small herbivorous mammal (25 g)
a
Application rate
(kg/ha)
FIR/bw a
Category
0.25
0.25
1.04
0.44
Small insects
Short grass
0.75
0.75
0.63
1.39
Large insects
Short grass
Mean (RUD)b
c
ftwa
(MAF)d
ETE e
TERflt
29
76
n/a
0.53
n/a
1.00
7.54
4.43
0.021
0.035
5.1
76
n/a
0.53
n/a
1.58
2.41
67.2
1.87
0.067
Food intake rate/body weight.
Residue per Unit Dose (residue after spraying normalised to application rate of 1 kg ai/ha).
c
Time-weighted average factor.
d
Multiple Application Factor.
e
ETE (mg/kg BW/d)=(FIR/bw)*RUD*MAF*ftwa. Other factors used to calculate the ETE (Crocker, 2005) are all assumed to be
equal to one (see text for details).
f
First-tier toxicity exposure ratio. Toxicity data used to calculate the TER were (i) the lowest NOEL in the avian reproduction test
(0.157 mg/kg BW/24 h) for insecticide I (Table 1); (ii) the two-generation rat toxicity NOEL (4.5 mg/kg BW/24 h) for fungicide F
(Table 2).
b
882 Shore et al.
TERs were calculated as the next step in the case
studies. It was assumed that onset of each reproductive phase occurred on the day after the compound was applied (п¬Ѓrst of the two applications for
compound F) and so each TER represented a worst
case scenario. The TERs were calculated for the
default indicator species used in п¬Ѓrst-tier assessments (herbivorous and insectivorous bird and
mammal) and mostly used their default values
(Table 3) when calculating the ETE term. However,
the default (21-day) time period used to calculate the
TWA factor (ftwa) (Table 3) was replaced by a time
period appropriate to the specific reproductive
phase (Bennett et al., 2005). Furthermore, the EU
guidance does not provide default ftwa or multiple
application factor (MAF) values (Table 3) when
calculating residues on insects because of a lack of
empirical data and difficulties in making theoretical
predictions. The default is to assume that there is no
decay of residues on insects (European Commission, 2002). In the case study with compound I,
however, data were available for rate of decay of
residues on insects (DT50 = 5.66) and this was
used to calculate TWAs for the insectivore phasespecific TERs. An example ETE calculation is given
in Box 1. The toxicity values were also adjusted
when calculating the phase-specific TERs. This involved using an extrapolation factor to incorporate
species differences in sensitivity to the pesticide
(Luttik et al., 2005). Unadjusted and adjusted
TERs, calculated using unadjusted and adjusted
NOELs, respectively, are both presented for the
avian (Table 4) and mammalian (Table 5) case
studies.
In the avian case study, all phase-specific,
unadjusted TERs were less than п¬Ѓve (Table 4).
This indicated that any reproductive phase might
be adversely affected if application occurred on the
day before onset of that phase. This assessment
provides clearer evidence than the п¬Ѓrst-tier TERlt
evaluation (Table 3) of the risk associated with the
proposed use of the compound. Furthermore, the
application of an appropriate exposure duration to
each toxicity endpoint means the TERs can effectively be used as a crude ranking system to identify
which reproductive phase may be most sensitive to
the pesticide; in this case study, the lowest unadjusted TERs were associated with copulation and
egg-laying (Table 4). One other noticeable difference between the first-tier and phase-specific
assessments was that the п¬Ѓrst-tier ETE was almost
twice as high for the insectivore as the herbivore
(Table 3) whereas the phase-specific ETEs were
similar for both indicators, but always a little
higher for the herbivore (Table 4). This difference
was largely because of the ways in which residues
on forage were estimated. First-tier assessments
tend to overemphasise the exposure of insectivores
relative to herbivores because of their assumption
of no residue decay on insects and use of a TWA
Box 1. Example calculation of ETE for a generic insectivorous bird in the course of egg formation at the time of spraying (see Table 4). Default values are from Table 3.
Equation
ETE = (FIR/BW) Г‚ C Г‚ AV Г‚ PT Г‚ PD (mg/kg BW/d)
where
C = C0 Г‚ MAF Г‚ ftwa
C0 = Application Rate (kg) Г‚ RUD (residue (mg/kg) per kg dose)
ftw = (I-e)kt)/kt
k = ln(2)/DT50
Specific values
t = 3 days
DT50 = 5.66
FIR/BW = 1.04
Calculations
1. k = ln(2)/5.66 = 0.122
2. kt = 0.122 Г‚ 3 = 0.367
3. C0 = 0.25 Г‚ 29 = 7.25
4. ftwa = (1)e)0.367)/0.367 = 0.837
5. C = 7.25 Г‚ 0.837 = 6.068
ETE = 1.04 Г‚ 6.068 = 6.311
Default values
Application rate = 0.25
RUD = 29
Assumptions
AV = 1 (no avoidance)
PT = 1 (all of diet contaminated)
PD = 1 (only insects eaten)
MAF = 1 (only 1 application)
Phase-specific effect of concern
Reduced fledgling survival from
direct exposure
Reduced juvenile survival and
growth from in ovo exposure
Adult behavioural effects leading to
abandonment of nesting attempt
Embryotoxicity leading to reduced
hatchability
Adult behavioural effects leading to
brood abandonment/abnormal
parental care
Reduced juvenile survival from
direct exposure
Reduced juvenile survival and
growth from in ovo exposure
5-day dietary toxicity
test with juvenilesf
Proportion of 14-day
juveniles per no/hatchlings
per hen
5-day dietary toxicity test
with juveniles
14-day-old juvenile
weights/hen
0.261
Mean eggshell
thickness/hen
Proportion of fertile
eggs/eggs set/hen
Change in adult BW
before egg laying
Proportion
hatching/eggs set/hen
Change in adult BW
before egg laying
0.289
0.067
0.261
0.029
0.157
9.53
0.289
9.53
0.608
0.067
0.261
20
0.608
0.067
0.067
0.029
0.608
ovum dvt TWA
5-day TWA
ovum dvt TWA
5-day TWA
2-day TWA
ovum dvt TWAd
1-day ETE
1-day ETE
1-day ETE
1-day ETE
1-day ETE
Adjusted Duration of
NOELb exposure
20
0.261
0.157
20
a
NOEL
No/eggs laid/hen
Change in adult BW
prior to egg laying1
NOEL test endpoint
used as surrogate
ETE
Insectivore
Herbivore
Insectivore
Herbivore
Insectivore
Herbivore
Insectivore
Herbivore
Insectivore
Herbivore
Insectivore
Herbivore
Insectivore
Herbivore
Insectivore
Herbivore
Insectivore
Herbivore
c
5.46
6.37
6.31
6.81
5.46
6.37
6.31
6.81
7.54
8.36
7.54
8.36
7.54
8.36
6.31e
6.81
6.60
7.04
Insectivore 7.54
Herbivore 8.36
Insectivore 7.54
Herbivore 8.36
Indicator
species
0.081
0.073
Adjusted
TER
1.75
1.50
0.041
0.038
1.75
1.50
0.025
0.023
0.035
0.031
0.035
0.031
2.65
2.39
0.041
0.038
3.03
2.84
0.053
0.045
0.011
0.010
0.053
0.045
0.005
0.004
0.009
0.008
0.009
0.008
0.081
0.073
0.011
0.010
0.092
0.086
0.021 0.004
0.019 0.003
2.65
2.39
TER
No observed effect level (NOELs) for dietary intake (mg/kg BW/d) calculated from no observed effect concentrations (NOECs) using data on the feeding rates and body
weights of test animals. Lowest value used when data for more than one species given in Table 1.
b
Toxicity factors adjusted to incorporate inter-species extrapolation using factors following Luttik et al. (2005).
c
ETE in mg/kg BW/d; case study data available for residues on foliage, so default DT50 of 10 days was used to calculate TWA factors.
d
Time weighted average (TWA) for an exposure period equivalent to length of time for an ovum to develop (ovum dvt) in species of interest (3 days in this calculation) –
primary in ovo exposure assumed to be from material deposited in the yolk during ovum formation.
e
The working for this calculation is shown in Box 1.
f
NOEC estimated as equivalent to LC05 and LC05 extrapolated from LC50 (Table 1).
a
Post-fledging survival
Juvenile growth
and survival
until fledging
Incubation and
hatching
Reduced fertility
Adult behavioural effects leading
to territory abandonment/delayed
breeding
Copulation and egg laying Adult behavioural effects leading
(5 days pre-laying through to reduced clutch size/abandonment
end of laying)
of nesting attempt
Reduced eggshell quality
Pair formation/breeding
site selection
Breeding phase
Table 4. Phase-specific avian toxicity data for insecticide I. Endpoints from Bennett et al. (2005) and toxicity values from Table 1
Long-Term Toxicity to Exposure Ratios (TERs) for Birds and Mammals 883
Adult behavioural effects leading to
abnormal litter care
Reduced litter survival
Increased developmental abnormalities
4.5
Behavioural observations
in 2-generation test
No/pups per mated female
in 2-generation test
Teratological effects
(2-generation or prenatal
development test)
28-day or 90-day toxicity tests
1.51
1.81
0.302
2.01
90
27
4.5
30
Insectivore
Herbivore
Insectivore
Herbivore
Insectivore
Herbivore
Insectivore
Herbivore
2.41
127
2.41
127
2.41
127
2.41
127
c
ETE
9.29
0.176
1.87
0.035
37.3
0.709
12.4
0.236
TER
1-day ETE
1-day ETE
Insectivore
2.41
Herbivore 127
Insectivore
2.41
Herbivore 127
1.87
0.035
12.4
0.236
Insectivore
2.41
9.29
Herbivore 127
0.176
1-day ETE
Insectivore
2.41 37.3
Herbivore 127
0.709
14-day TWA Insectivore
2.41d 37.3
Herbivore 81.1
1.11
14-day TWA Insectivore
2.41d 11.2
Herbivore 81.1
0.333
1-day ETE
1-day ETE
1-day ETE
1-day ETE
1-day ETE
Indicator
species
0.125
0.002
0.834
0.016
0.627
0.012
2.61
0.050
0.627
0.019
0.751
0.022
0.627
0.012
0.125
0.002
2.61
0.050
0.834
0.016
Adjusted
TER
No observed effect level (NOELs) for dietary intake (mg/kg BW/d) either as reported or calculated from no observed effect concentrations (NOECs) using data on the feeding
rates and body weights of test animals.
b
Toxicity factors adjusted to incorporate inter-species extrapolation using factors following Luttik et al. (2005).
c
ETE in mg/kg BW/d.
d
Data not available for residues on insects and European Commission (2002) guideline (assumes no decay over time in residue concentrations in insects) followed when
calculating time-weighted ETE.
a
6.30
1.51
2.01
6.30
0.302
1.51
Adjusted Duration of
NOELb exposure
90
22.4
30
90
22.4
a
NOEL
28- or 90-day toxicity tests
NOEL test endpoint used
as surrogate
No/weanlings per mated female
in 2-generation test
Post-weaning survival Reduced juvenile survival
Survival to 4 weeks in
until maturity
2-generation test
Reduced juvenile growth and development Body weight of 4-week-old
juveniles
in 2-generation test
Reproduction of F1 Reduced productivity of F1 generation
No/weanlings per mated F1
generation
female in 2-generation test
Increased developmental
Teratological effects in
abnormalities in F2
2-generation test
Pup growth and
survival
until weaning
Adult behavioural effects leading to
territory abandonment or delayed or
abnormal mating
Establishing
breeding site,
pairing, and
mating
Pregnancy
Reduced litter size
Phase-specific effect of concern
Breeding phase
Table 5. Phase-specific mammalian toxicity data for fungicide F. Endpoints from Bennett et al. (2005) and toxicity values from Table 2
884 Shore et al.
Long-Term Toxicity to Exposure Ratios (TERs) for Birds and Mammals 885
for residues on foliage. The phase-specific ETEs
adopted a more equitable approach by using either
a one-day exposure or relatively short TWAs for
residues on both insects and foliage.
In the mammalian case study, the phase-specific
TERs (Table 5) generally confirmed the first-tier
assessment for compound F in that many of the
TERs were lower than the trigger value. The level
of risk indicated for the herbivore was higher in
the phase-specific than in the first-tier assessment
because the exposure associated with the toxicity
endpoint was generally higher. As in the avian case
study, the fact that all phase-specific TERs were
below п¬Ѓve for the herbivore provides stronger
grounds, compared with a single п¬Ѓrst-tier TERlt,
for requiring additional information prior to authorisation of the product. Unlike the herbivore,
the insectivore unadjusted, phase-specific TERs
fell below the trigger value only on two reproductive phases, establishment of breeding site and
productivity of the F1 generation (Table 5). The
TERs for both phases were based on a worst case
(onset of phase the day after spraying) one-day
ETE. This suggests that further refinement of the
exposure term, or alteration of spraying date so
that it does not coincide with onset of breeding,
may be sufficient to reduce risk to insectivores to
an acceptable level.
Incorporation of an inter-species extrapolation
factor in the risk assessment (by using adjusted
NOELs) reduced avian (Table 4) and mammalian
(Table 5) phase-specific TERs by 4–33-fold and
14–60-fold, respectively. Bennett et al. (2005)
proposed that the uncertainty factor of п¬Ѓve applied
to the п¬Ѓrst-tier TERlt should be reduced to one
when using adjusted TERs because the main purpose of the uncertainty factor was to allow for the
possibility that some wildlife species may be more
sensitive to the pesticide than the test species. The
present case studies indicated that, if the extrapolation factors for toxicity values proposed by
Luttik et al. (2005) are accepted, a п¬Ѓvefold uncertainty factor is often likely to be inadequate for
allowing for variation in species sensitivity.
Therefore, the overall effect of using adjusted
NOELs was that the phase-specific TERs for
pesticides I and F more often fell below their
trigger value.
Phase-specific long-term TERs for skylark
and wood mouse
The exposure data used to calculate the phasespecific TERs in Tables 4 and 5 applied generically
to any herbivore or insectivore of a given weight
and assumed that diet was monophagous and always contaminated (PD and PT were both 1).
Thus, they again were relatively worst-case scenarios that did not account for factors that might
moderate the risk to species actually present in
cereals. The next stage of the case studies was to
calculate phase-specific TERs for the skylark and
wood mouse, two species potentially at high risk of
pesticide exposure in cereal п¬Ѓelds in Britain. The
ways in which the ETE component of the TER can
be refined for particular species is discussed by
Crocker (2005). Relevant ecological data for the
skylark and the wood mouse are summarised by
Roelofs et al. (2005) and were used here to refine
the ETE for compounds I and F, respectively.
When refining the ETE, it was assumed that
exposure to the pesticide was exclusively via the
diet and not through skin contact or inhalation,
although such routes may be important under п¬Ѓeld
conditions (Mineau, 2002). It was also assumed
that animals were exposed to pesticides in proportion to the amount of time they spent in habitat
where pesticides are used, but did not distinguish
between crops or allow fields to receive different
spray regimes. Other assumptions were that diet
composition was independent of habitat use,
individuals ate the same diet on successive days,
and animals did not show any avoidance of pesticide residues, although this can occur to some
extent under п¬Ѓeld conditions (McKay et al., 1999).
Daily food intake was estimated from information
about daily energy expenditure and energy and
moisture content of wildlife food (Crocker et al.,
2002a). The total daily intake of pesticide (mg/kg
BW/day) was calculated as shown in Box 2. For
each of the parameters shown, wherever empirical
data were available, calculations were based on the
mean values, i.e. we did not attempt to build in
worst-case assumptions by using 95th or other
percentiles.
The refined ETEs for the skylark (Table 6) were
nearly always higher than those for a generic bird
886 Shore et al.
Box 2. Example calculation of ETE for an adult skylark in the course of egg formation at the time of spraying (see Table 5).
Equations
FIR
Г‚ AV Г‚ PT
BWt
DEE
Г°Ci Г‚ PDiГћ Гѕ Г°Cs Г‚ PDsГћ Гѕ Г°Ch Г‚ PDhГћ
ETE Вј
Г‚
Г‚ AV Г‚ PT
BWt
Г° PDi Г‚ Г°1 ГЂ MiГћ Г‚ GEi Г‚ AEiГћ Гѕ Г°PDs Г‚ Г°1 ГЂ MhГћ Г‚ GEs Г‚ AEsГћ Гѕ Г° PDh Г‚ Г°1 ГЂ MhГћ Г‚ GEh Г‚ AEhГћ
where:
i=insects, s=seeds, and h=herbs
FIR=food intake rate (g fresh wt/day)
DEE=Daily Energy Expenditure (kJ)
BWt=Body Weight (g)
C=Concentration (mg ai/kg wt) of pesticide residue on food
PD=Proportion (fresh wt) of Diet made up by food type (0–1)
M=Proportion Moisture in fresh food (0–1)
GE=Gross Energy (kJ/g dry wt) provided by food type
AE=Assimilation Efficiency of food type (0–1)
PT=Proportion of food eaten that is from the treated area (0–1)
AV=Avoidance of treated food (0=wholly avoided, 1=not avoided)
Log(DEE) = 1.0017 + 0.7034 Г‚ log(BWt )
C=C0 x MAF Г‚ ftwa
ftwa=(I)e)kt)/kt
k=ln(2)/DT50
ETE Вј
Assumptions
AV=1 (no avoidance)
PT=0.51 of diet contaminated
MAF=1 (only 1 application)
Calculations
Specific values
Food
B Wt
g
DEE
kJ
Insects
Seeds
Herbs
37.2
37.2
37.2
128
128
128
Calculated values
PD
GE
kJ/g
dry wt
M
AE
C0
mg/kg
0.300
0.200
0.500
22.7
20.0
17.1
0.660
0.125
0.779
0.76
0.80
0.58
7.50
25.0
60.0
(Table 4). This was despite PT being reduced from
the п¬Ѓrst-tier default of one to 0.51, a value derived
from radio-tracking studies on skylarks in arable
fields in Britain (Crocker – unpub. data). Some
50% of the adult skylark diet in spring is comprised of leaves (Green, 1978), and the foliar RUD
for compound I, estimated from empirical data
(Mineau et al., 2001), was approximately three
times the п¬Ѓrst-tier default value. This high foliar
RUD largely explained why adult skylark ETEs
were usually greater than those for a generic herbivorous or insectivorous bird. Nestling skylarks
are wholly insectivorous (Donald, 2004), yet the
phase-specific ETEs based solely on direct dietary
exposure of juveniles were lower, not higher, than
the default value for a generic insectivore. This was
partly due to the fact that only half of the diet was
t
days
3
3
3
DT50
days
FIR
PT g/day
k
ftwa
5.66
5.06
10.0
0.51 6.82
0.51 4.54
0.51 11.4
0.12
0.14
0.07
0.84
0.82
0.90
C
mg/kg
ETE
mg/kg
6.28
20.5
54.2
Sum
0.59
1.28
8.44
10.31
assumed to be contaminated (PT=0.51). Another
factor was that their daily energy expenditure
(DEE) was estimated in a different way from the
default п¬Ѓrst-tier calculation. The DEE of inactive
nestlings kept warm by a brooding mother is unlikely to be a good measure of food intake. A large
part of nestling food intake will go towards weight
gain. Therefore, food intake in nestlings was based
on empirical data for passerine chick feeding rates.
Although it was assumed that, once out of the
nest, fledgling skylarks (estimated BW 30 g),
would follow the standard allometric equation
(second equation in Box 2), heavier birds eat less in
proportion to body weight than smaller birds.
Hence it was predicted that exposure would also
be lower for a 30 g fledgling skylark than for a
10 g generic passerine.
28- or 90-day toxicity tests
Behavioural observations in 2-generation test
No/pups per mated female in 2-generation test
Teratological effects in 2-generation test or prenatal
development toxicity test
28-day or 90-day toxicity tests
No/weanlings per mated female in 2-generation test
Survival to 4 weeks in 2-generation test
Body weight of 4-week-old juveniles in 2-generation test
No/weanlings per mated F1 female in 2-generation test
Teratological effects in 2-generation test
Change in adult BW prior to egg laying
No/eggs laid/hen
Mean eggshell thickness/hen
Proportion of fertile eggs/eggs set/hen
Change in adult BW before egg laying
Proportion hatching/eggs set/hen
Change in adult BW before egg laying
5-day dietary toxicity test with juveniles
Proportion of 14-day juveniles per
no/hatchlings per hen
5-day dietary toxicity test with juveniles
14-day-old juvenile weights per hen
NOEL used as surrogate
No observed effect level (NOELs) and ETE values expressed as mg/kg BW/d.
Reproduction of F1 generation
Post-weaning survival until maturity
Pup growth and survival until weaning
Pregnancy
Wood mouse
Establishing breeding site, pairing, and mating
Post-fledging survival
Juvenile growth and survival until fledging
Incubation and hatching
Skylark
Pair formation/breeding site selection
Copulation and egg laying
Breeding phase
22.4
90
90
27
4.5
30
22.4
4.5
90
30
9.53
0.261
20
0.157
0.261
0.261
20
0.261
20
9.53
0.157
NOEL
1.51
6.30
1.51
1.81
0.302
2.01
1.51
0.302
6.3
2.01
0.289
0.067
0.608
0.029
0.067
0.067
0.608
0.067
0.608
0.289
0.029
Adjusted
NOEL
7.91
7.91
5.17
5.17
7.91
7.91
7.91
7.91
7.91
7.91
1.81
10.3
11.6
11.6
11.6
11.6
11.6
10.3
10.7
1.78
10.3
Species-specific
ETE
2.83
11.4
17.4
5.22
0.569
3.79
2.83
0.569
11.4
3.79
5.26
0.025
1.72
0.014
0.022
0.022
1.72
0.025
1.87
5.35
0.015
TER
0.191
0.796
0.292
0.350
0.038
0.254
0.191
0.038
0.796
0.254
0.160
0.007
0.052
0.002
0.006
0.006
0.052
0.007
0.057
0.162
0.003
Adjusted
TER
Table 6. Phase-specific TERs calculated for the skylark and the wood mouse. Onset of each reproductive phase assumed to occur on the day after the application of the
pesticide. Duration of exposure for each reproductive phase is indicated in Tables 4 and 5 and the species specific data used to refine components of the ETE are given in
Roelofs et al. (2005).
Long-Term Toxicity to Exposure Ratios (TERs) for Birds and Mammals 887
888 Shore et al.
For the wood mouse, PT was again reduced
(this time from 1 to 0.38) on the basis of radiotracking data (Crocker et al., 2002b). Arable wood
mice eat seeds, insects and plant material (Watts,
1968; Gorman and Zubaid, 1993; Rogers and
Gorman, 1995) and the estimated residue of
compound F for this omnivorous diet (Roelofs
et al., 2005) falls between those based on default
RUDs for an obligate herbivore and obligate
insectivore (Table 3). This largely accounted for
why the phase-specific ETEs for the wood mouse
(Table 6) were always lower than the default values for a generic herbivorous mammal but similar,
although somewhat higher, than those for a generic insectivore (Table 5).
The effect of using species-specific ETEs on the
TER calculation differed between the case studies.
Because almost all the ETE values increased
when they were refined for the skylark, the
associated phase-specific TERs (Table 6) were
mostly smaller than those for a generic herbivorous or insectivorous bird (Table 4), and were
below п¬Ѓve in each case. The only exceptions were
those associated with the direct toxicity of compound I to the chick or fledgling. In contrast, six
of the 10 refined TERs for the wood mouse had
values that exceeded п¬Ѓve and another two values
exceeded three (Table 6); most of the generic
mammalian insectivore TERs but none of the
generic mammalian herbivore TERs exceeded п¬Ѓve
(Table 5). Thus, on the basis of the refinement of
the ETE, the predicted direct toxicity of compound F to the wood mouse would appear to less
than that predicted for a generic mammalian
herbivore such as a vole, but largely similar to
that predicted for a generic mammalian insectivore such as a shrew. However, it is arguable that
if a species-specific ETE is used to calculate the
TER, the appropriate toxicity endpoint is an
avian or mammalian NOEL that is adjusted to
account for inter-species variation in sensitivity.
None of the adjusted TERs exceeded one, the
trigger value proposed by Bennett et al. (2005),
for the skylark while only two of the adjusted
TERs exceeded this value for the wood mouse
(Table 6). Use of species-specific ETEs and adjusted NOELs would, therefore, indicate that
compounds I and F could potentially have significant impacts on multiple reproductive phases
of the skylark and wood mouse, respectively.
Chronological TERs
The phase-specific TERs in Table 6 remain worstcase scenarios in that they assume spraying occurs
the day before onset of each reproductive phase. In
reality, breeding may begin before or well after
pesticide application and few, if any, of the specific
reproductive phases may occur at, or close to, the
spraying date. Thus, many of the ETE values in
Table 6 could be overestimated. To evaluate how
significant the temporal dynamic between spraying
date and onset of reproduction might be, limited
temporal variation was incorporated into the case
studies. This was done by assuming a spraying
date for both compounds of May 1st, which was
within the distribution of spraying times for
insecticides and fungicides in Britain (Roelofs
et al., 2005). Onset of the п¬Ѓrst reproductive phase
(Table 6) for skylarks and wood mice in arable
п¬Ѓelds was assumed to be two days after and
10 days before spraying, respectively. This was
consistent with data from empirical studies on the
timing of breeding for these species (Donald et al.,
2001; Tattersall and Macdonald, 2003). The relevant data on breeding biology that were used to
estimate the timing of onset of each subsequent
reproductive phase are given in Roelofs et al.
(2005).
The effect of incorporating a breeding calendar
and a spraying date of May 1st was to reduce
estimated exposure for most reproductive phases
(Table 7). For the skylark, the ETEs for each
phase were approximately 70% of the values calculated in Table 6, although the reduction was
substantially greater in some cases. The phasespecific TERs were elevated when they were based
on the revised ETEs (Table 7). Despite this, the
overall assessment of risk was the same as when no
calendar was applied (Table 6), in that only the
unadjusted TERs for 5-day toxicity to chick and
fledgling exceeded their trigger value of five.
However, in both these cases, the adjusted TERs
also exceeded their trigger value of one (Table 7).
When the calendar was applied in the wood mouse
case study, the effects on the ETE were marked
(Table 7). Exposure during the п¬Ѓrst reproductive
phase was zero because this phase preceded the
spraying date. The ETEs during subsequent
reproductive phases were reduced to an average of
29% (range 1–66%) of their values in Table 6,
Body weight of 4-week-old juveniles in
2-generation test
No/weanlings per mated F1 female in
2-generation test
Teratological effects in 2-generation test
28- or 90-day toxicity tests
Behavioural observations in 2-generation test
No/pups per mated female in 2-generation test
Teratological effects in 2-generation test or
prenatal development toxicity test
28-day or 90-day toxicity tests
No/weanlings per mated female in
2-generation test
Survival to 4 weeks in 2-generation test
5
5
5
11
TWA
TWA
TWA
TWA
No/eggs laid/hen
Mean eggshell thickness/hen
Proportion of fertile eggs/eggs set/hen
Change in adult BW
Proportion hatching/eggs set/hen
Change in adult BW
5-day dietary toxicity test with juveniles
Proportion of 14-day juveniles per
no/hatchlings per hen
5-day dietary toxicity test with juvenilesb
14-day-old juvenile weights per hen
over
over
over
over
days
days
days
days
5–7
22–23
22–26
5–7
Day 65 after 2nd spray
14 day TWA beginning
Day 24 after second spray
Day 6 after 2nd spray
)10 to )7
)10 to )7
Day 0 of 2nd spray
Day 0 of 2nd spray
TWA over days 40–44
TWA over days 5–7
2
Days after sprayinga
Change in adult BW prior to egg laying
NOEL used as surrogate
50.6
337
0.089
28.0
93.4
4.29
17.2
NA
NA
11.4
3.79
705
0.036
0.020
0.034
0.034
4.09
0.038
9.96
79.1
0.022
2.03
TER
0.089
0.964
0.964
5.22
5.22
0
0
7.91
7.91
0.014
6.93
7.77
7.77
7.77
4.89
6.93
2.01
0.120
6.93
9.87
Species-specific
ETE (mg/kg BW/d)
22.6
3.39
1.88
1.57
0.289
1.21
NA
NA
0.796
0.254
21.4
0.009
0.004
0.009
0.009
0.124
0.010
0.30
2.40
0.004
0.062
Adjusted
TER
a
For the skylark, the breeding cycle was assumed to follow this calendar: arrival and pairing, May 3–5; copulation May 6; egg development (3 days) May 6–12, Egg-laying (4
eggs) May 9–12; incubation May12–22; nestling May 23–30, post-nestling May 31–June 9; fledging 10 June. For the wood mouse the breeding calendar was: arrival and pairing
April 21–24, copulation April 25, pregnancy April 26–13 May; lactation 14–31 May; post-weaning 1 June–11 July; mature 12 July.
b
Juveniles assumed to weigh 30 g, with DEE of 110 kJ (see Box 2), and wholly insectivorous.
Reproduction of
F1 generation
Post-weaning survival until
maturity
Pup growth and survival
until weaning
Wood mouse
Establishing breeding site,
pairing, and mating
Pregnancy
Post-fledging survival
Juvenile growth and survival
until fledging
Incubation and hatching
Skylark
Pair formation/breeding
site selection
Copulation and egg laying
Breeding phase
Table 7. Phase-specific TERs calculated for the skylark and the wood mouse assuming that first application of pesticide is on May 1st and timing of reproduction occurs
as outlined by Roelofs et al. (2005). Duration of exposure for each reproductive phase and the adjusted and non-adjusted NOEL values used to calculate the TER are
indicated in Tables 4 and 5
Long-Term Toxicity to Exposure Ratios (TERs) for Birds and Mammals 889
890 Shore et al.
except during pregnancy when exposure was
maximal because it coincided with the second
fungicide spray. The associated effects on the
various TERs for wood mice were that eight
unadjusted values were greater than п¬Ѓve and the
remaining two were higher than three. The magnitude of the reduction in estimated exposure in
wood mice was such that all but three of the
adjusted TERs also exceeded their trigger value. It
is clear from these two case studies that the effects
of applying a calendar to the phase-specific toxicity endpoints can be pronounced, but vary substantially for different compounds or species.
Discussion
This paper used case studies to demonstrate how
deterministic long-term risk assessments can be
enhanced by use of multiple toxicity endpoints and
refined estimates of exposure. The studies progressed from the current EU first-tier assessment
to incorporate phase-specific end-points, adjustment of the ETE so that it applied to specific
(skylark, wood mouse) rather than generic indicator species, and then further refinement of the
ETE so that timing of pesticide application was
related to onset of breeding. The other major aspect incorporated into the case studies was the
replacement of unadjusted NOELs with adjusted
values that accounted for species variation in
sensitivity to toxic insult. Although there were
only two case studies, they provided evidence of
some of the major impacts these modifications can
have on assessment of risk.
In theory, the use of multiple rather than single
toxicity endpoints should not per se affect the
outcome of a long-term risk assessment. This is
because the current п¬Ѓrst-tier TERlt uses the lowest
of the various NOELs in the phase-specific
assessments. The only exception is the use of the 5day dietary toxicity test to assess impacts on
growth and survival of chicks and fledglings. This
is not assessed in the avian reproduction test, because chicks are reared on uncontaminated diet
(Mineau, 2005). Thus, a compound with high
dietary toxicity to chicks could fall below the
trigger value for a phase-specific assessment but
exceed the trigger value for a standard п¬Ѓrst-tier
TERlt. In practice however, use of phase-specific
endpoints requires concurrent reduction of the
time period used in the ftwa from the default 21days to values that range from one to 14 days
(Tables 4 and 5). The resultant ftwa will normally
be greater than the default value unless there is
concentration, or at least no degradation, of dietary residues over time. Typically therefore, the
lowest of the NOELs in the phase-specific assessments will be evaluated against a higher ETE than
in the current п¬Ѓrst tier assessments, although the
difference in the case studies was, at most, only
twofold. Consequently, the phase-specific risk
assessment is likely to be somewhat more stringent
than the current п¬Ѓrst-tier assessment if there is no
subsequent further refinement of the ETE.
Adjustment of the exposure term so that it applies to individual species, rather than generic
indicators, might have been expected to reduce the
ETE values in the case studies, not least because
PT was decreased by between a half and twothirds. However, it was apparent that replacing
generic default values with species-specific exposure parameters can have complex impacts and, in
some cases, negate the influence of even a relatively large reduction in PT. The refined ETE for
the skylark was, in fact, higher than that for the
generic herbivorous and insectivorous bird (Tables 4 versus 6), demonstrating that the п¬Ѓrst-tier
assessment is not always particularly conservative.
Although omnivorous, skylark diets can contain
substantial amounts of leaf material, and being
relatively small birds, they may be more at risk
than larger obligate herbivores. In contrast, the
mammal case study indicated that the estimated
exposure for wood mice was similar to that predicted for the generic insectivore (shrew) but much
lower than that predicted for the generic herbivorous vole (Tables 5 versus 6). Intuitively, it might
be thought reasonable to use the generic vole (but
not the shrew) as a surrogate for wood mice as
both are rodents, have similar body weights and
some dietary overlap (Hansson, 1985; Corbet and
Harris, 1991). However, the case study data suggests that if this is done, exposure in wood mice is
markedly overestimated.
In reality, most of the ETE values calculated
without the implementation of a calendar are
likely to be overestimated because they are assumed to be close to, or at, a maximum during
each reproductive phase. Although the reduction
Long-Term Toxicity to Exposure Ratios (TERs) for Birds and Mammals 891
in ETE estimates associated with use of a breeding
calendar and spraying date did not raise many
skylark TERs above the trigger value, most wood
mouse chronological TERs did exceed the trigger
value. The only exceptions were for effects on
pregnancy and direct toxicity to pups and were due
to a high level of exposure following the second
fungicide application (Table 7). The chronological
TERs therefore indicated a potential to reduce the
predicted risk by limiting fungicide use to a single
application. The simple п¬Ѓrst-tier TERlt values
(Table 3) do not provide such insight. Even when
the TERlt values were recalculated assuming a
single fungicide application, the resultant values
(1.87 for the generic insectivore, 0.107 for the small
herbivore) fell below the trigger value.
Although chronological TERs may provide a
more realistic assessment of risk, they are deterministic values and, as such, problematic. This is
because they are based on one single specific date
for pesticide application and another for uniform
onset and procession of breeding. The calculated
TERs are likely to vary markedly with choice of
dates. For example, the coincidence of the second
fungicide spraying with pregnancy in the wood
mouse case study could disappear if a different
calendar was used. In reality, spraying date(s) and
the dates for onset of each reproductive phase all
vary. Chronological phase-specific assessments
therefore demand a probabilistic approach that
incorporates information about the variation in
these parameters and how they interact with each
other. Such an approach would also allow incorporation of data on the distributions of other
parameters in the model (food consumption, energy expenditure, pesticide residues, species sensitivity, etc), all of which also vary, although to
differing degrees and not all contribute equally to
uncertainty in the risk assessment. The type of
model needed to produce this type of probabilistic
assessment and the nature of the resultant outputs
are beyond the scope of this paper but are considered by Roelofs et al. (2005).
The п¬Ѓnal major aspect of the long-term risk
assessment that was investigated in the case
studies was to examine the effect of using adjusted rather than unadjusted NOELs to calculate
TERs. Adjusted NOELs are usually, but not always, lower than unadjusted values (Luttik et al.,
2005). This was true in the case studies, and
replacement of unadjusted NOELs with adjusted
values increased the proportion of phase-specific
TERs that fell below the trigger value, even
though this was relaxed from п¬Ѓve to one. Thus,
the uncertainty factor of п¬Ѓve applied to the
unadjusted TERlt would seem insufficient to account for species variation in sensitivity to pesticides, let alone other sources of uncertainty. This
suggests again that the current п¬Ѓrst-tier TERlt
assessment may not be sufficiently protective, but
in this case because toxicity, rather than exposure, is underestimated.
Table 8. Decision tree for assessing the long-term risk to vertebrates from novel compounds
Criteria if NOELs are
Stage
Description
Unadjusted
1
Standard п¬Ѓrst tier TERlt
2
Phase specific TERs for generic indicator speciesa
3
Phase specific TERs for particular indicator speciesb
4
Deterministic chronological TERsc
TERlt
TERlt
TERlt
TERlt
TERlt
TERlt
TERlt
TERlt
5
Probabilistic assessment of risk and potential population effectsd
a
‡5
<5
‡5
<5
‡5
<5
‡5
<5
Adjusted
TERlt
TERlt
TERlt
TERlt
TERlt
TERlt
TERlt
TERlt
‡1
<1
‡1
<1
‡1
<1
‡1
<1
Next stage
Stop
To stage 2
Stop
Stage 3
Stop
Stage 4
Stop
Stage 5
Calculated using TWA factor appropriate to the specific toxicity endpoint.
If skylark and wood mouse are used as specific indicators, the required ecological data to refine the ETE are summarised by Roelofs
et al. (2005).
c
Calculated using a single spraying date consistent with the proposed use of the compound and a single date for onset of breeding for
indicator species.
d
As described by Roelofs et al. (2005).
b
892 Shore et al.
The original brief of the risk assessment
workshop was to remain close to the current EU
guidance for long-term risk assessment. It is
therefore pertinent to ask how the approaches
outlined here can be readily implemented. A
decision tree of rules was suggested (Table 8) that
followed the assessment sequence outlined in the
present paper. It was proposed that phase-specific
assessments should be carried out only for compounds that did not have п¬Ѓrst-tier TERlt values
of п¬Ѓve or more. This conforms to current EU
guidance but the case studies have shown that
the п¬Ѓrst-tier assessment may under-estimate risk
in some cases. One solution would be to increase
the conservatism of standard п¬Ѓrst-tier assessments. Alternatively, it might be argued that all
compounds should be subject to both the п¬Ѓrsttier TERlt and the subsequent assessment stages
in Table 8, thereby making more use of the
available information. Unadjusted or adjusted
NOELs (following Luttik et al., 2005) could be
used. The procession from deterministic chronological TERs to full probabilistic assessment
(Stage 4 to 5 in Table 8) is a large and complex
step and application of uncertainty bounds
analysis to the chronological TERs may be
merited п¬Ѓrst. This would assess the likely range
of variation in the chronological TER values and
help determine if a full probabilistic assessment
was necessary.
In conclusion, this paper has shown how the
toxicity data required under current registration
procedures can be used to produce more realistic
assessments of long-term risk to birds and mammals. Implementation of staged refinements to the
deterministic assessment is currently possible using
existing registration data for the compound and
available ecological data for indicator species. The
advantages of using such phase-specific assessments are that they potentially give a more accurate assessment of risk and identify the most
vulnerable phase(s) of reproduction. This information can further be used to formulate riskreduction measures, as demonstrated in the wood
mouse case study, or to target data-gathering
needed for refining the assessment. Phase-specific
assessments are also a necessary prerequisite for
developing methods to predict impacts at the
population level.
Acknowledgments
We acknowledge the UK Department for Food
and Rural Affairs and its Pesticide Safety Directorate for funding the workshop on long-term
risk assessment.
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