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SID 5
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Research Project Final Report
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•
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Project identification
1.
Defra Project code
2.
Project title
PS2302
Acceptability of Pesticide Effects on Non-Target Species
3.
Contractor
organisation(s)
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Crane Consultants
Horticulture Research International
Virginia Insitute of Marine Sciences
University of Warwick
University of Reading
University of Edinburgh
Imperial College
MORI Social Research Institute
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SID 5 (2/05)
4. Total Defra project costs
5. Project:
Page 1 of 30
£
143750.00
start date ................
01 November 2002
end date .................
31 October 2004
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Executive Summary
7.
The executive summary must not exceed 2 sides in total of A4 and should be understandable to the
intelligent non-scientist. It should cover the main objectives, methods and findings of the research, together
with any other significant events and options for new work.
EU Directive 91/414/EEC requires there to be no unacceptable effects on the environment from the use of pesticides.
This report examines the ecological and socioeconomic bases for establishing Ecologically Acceptable
Concentrations (EACs), with the following specific objectives:
1.
To produce a comprehensive literature review on current knowledge and understanding of the ecological
significance of long-term risks from pesticides to non-target species.
2.
To run focus groups with particular stakeholders and to perform a statistically robust public opinion survey to
determine the spectrum of views on EACs and the factors that influence these views.
3.
To combine information from 1 and 2 above within a formal decision analysis framework to provide PSD with a
coherent, logical, yet flexible decision tool for identifying EACs under Directive 91/414/EEC.
Review of the ecological literature showed that although much is known about the potential for pesticides to cause
long-term risks to non-target organisms, there is little information on the frequency or extent that these risks are
realised in the natural environment. Current risk assessment methods would allow reasonable predictions of longterm effects of pesticide application if three changes were instituted. First, more population-based laboratory studies
should be applied in predictive pesticide risk assessment. Second, risk assessment should include as much effort on
collating and integrating ecological knowledge into the assessment in Tier 1 as is currently expended on gathering
chemical and toxicological information on exposure and effects. Production of a formal conceptual ecological risk
assessment model for each product or active substance for which authorisation is sought would provide an
appropriate framework for integrating and applying such knowledge. Third, in acknowledgment of the uncertainties
in the predictive risk assessment process, more post-authorisation monitoring should be done.
The review clearly identified the existence of large uncertainties when extrapolating from laboratory tests to
environmental effects, and the continuing need for elements of judgement to be applied when authorising active
substances. The next stage of the project examined what stakeholder and public values should be taken into account
when these judgements are made. Stakeholders in focus groups were almost entirely concerned with the potential
effects of pesticides on animal and plant population viability, and micro-organism function. Stakeholders in focus
groups also recognised that a trade-off exists between the potential economic advantages of responsible pesticide use
and the potential disadvantages of individual poisoning events. A subsequent public opinion survey of 2000+
respondents showed that although pesticides are widely used in homes and gardens, their use on farm crops remains
of concern to the public. Concerns are greatest on issues of human health and food quality, but potential
SID 5 (2/05)
Page 2 of 30
environmental effects are also an issue for a substantial number of people, particularly if attractive species could be
affected. In contrast to the stakeholders in focus groups, substantial numbers of the public would remain very
concerned about the effects of pesticides even if they affected only individual organisms and not populations.
However, the desire to purchase food that has not been produced with the use of pesticides depends on its price
relative to the total available budget, with only one-fifth of the public, or fewer, prepared to pay a substantial
additional amount for this. Pesticide risk managers need to ensure that long-term adverse pesticide effects on
populations of plants and animals, or on microbial function, do not occur. This can be achieved by use of improved
predictive effects and exposure models combined with more effective post-authorisation monitoring. Individual
vertebrate deaths should be avoided, if at all possible, even if these deaths are unlikely to affect population size or
viability, and particularly if the organisms at risk are those that the public care about. However, a substantial
proportion of the public is unlikely to be satisfied by any risk mitigation for pesticides, even if evidence shows that it
results in negligible risks to the environment.
The information from earlier project components was used to construct a formal decision tree to assist risk managers
and their advisors in coming to consistent decisions on pesticide authorisations, informed by public values. Scientific
evidence takes precedence over opinion in this framework. However, public values are used when decisions are
value-laden rather than evidence-based.
Project Report to Defra
8.
As a guide this report should be no longer than 20 sides of A4. This report is to provide Defra with
details of the outputs of the research project for internal purposes; to meet the terms of the contract; and
to allow Defra to publish details of the outputs to meet Environmental Information Regulation or
Freedom of Information obligations. This short report to Defra does not preclude contractors from also
seeking to publish a full, formal scientific report/paper in an appropriate scientific or other
journal/publication. Indeed, Defra actively encourages such publications as part of the contract terms.
The report to Defra should include:
z the scientific objectives as set out in the contract;
z the extent to which the objectives set out in the contract have been met;
z details of methods used and the results obtained, including statistical analysis (if appropriate);
z a discussion of the results and their reliability;
z the main implications of the findings;
z possible future work; and
z any action resulting from the research (e.g. IP, Knowledge Transfer).
SID 5 (2/05)
Page 3 of 30
1. INTRODUCTION
EU Directive 91/414/EEC (EC 1993) harmonises the registration of pesticides throughout the European Union. The
environmental objectives of the Directive are contained in the preamble which states that:
‘…plant protection products can have non-beneficial effects upon plant production…[and]…their use may involve
risks and hazards for humans, animals and the environment, especially if placed on the market without having been officially
tested and authorised and if incorrectly used.
…the provisions governing authorisation must ensure a high standard of protection, which, in particular, must
prevent the authorisation of plant protection products whose risks to health, groundwater and the environment and human
and animal health should take priority over the objective of improving plant production.
….it is necessary, at the time when plant protection products are authorised, to make sure that, when properly
applied for the purpose intended, they are sufficiently effective and have no unacceptable effect on plants or plant products,
no unacceptable influence on the environment in general, and, in particular, no harmful effect on human or animal health or
on groundwater.’
'Animals' are defined in the Directive as 'belonging to species normally fed and kept or consumed by man,' so this
does not include wild flora or fauna. 'Environment' is defined as 'water, air, land and wild species of fauna and flora, and any
interrelationship between them, as well as any relationship with living organisms.'
A key phrase in the Directive, which is quoted above, is that pesticides should have 'no unacceptable influence on
the environment in general…' Crane and Giddings (2004) argue that the acceptability or otherwise of an environmental
influence (i.e., an adverse effect on the environment) clearly involves social values and should be based upon consultation
with a wide group of stakeholders, including environmental scientists, government regulators and pesticide manufacturers.
They also argue that representatives from the wider community and environmental Non-Governmental Organisations should
be consulted (Crowfoot and Wondolleck 1990). However, to date, the concept of the ‘acceptability’ of pesticide effects in
Europe, through definition of Ecologically Acceptable Concentrations (EACs), has been developed and refined largely by
scientists from the regulatory and business communities, who may not represent the full spectrum of views on what
constitutes an acceptable environmental effect. At one time scientists were seen as experts who could provide an unbiased
opinion on scientific matters, and whose advice would be accepted by decision makers on the basis of that acknowledged
expertise. The unbiased nature of this knowledge was generally accepted until some damaging decisions, particularly in the
agricultural sector, caused questions to be asked about the decision-making process and the nature of scientific evidence.
Research in the social sciences has also turned up complex relationships between scientific results and assessment, trust and
public perception (Douglas 2000).
With an increasingly sceptical society it is important that scientists and non-specialists communicate and, in
particular, discuss complex ideas. This is necessary because the public’s perception of risks may well diverge significantly
from that of specialists (Frewer 2004, Hansen et al. 2003). An individual’s perception of risk depends upon an often intuitive
judgement of the probability of occurrence and the severity of the consequences of that risk. This perception is usually a
judgement that is made without consideration of associated benefits, and risks only become acceptable to an individual when
they are able to balance them with these benefits. However, even if individuals agree on the degree of risk, they may still
disagree on its acceptability because of differences in their level of expertise and education, their gender or their personal
values. For example, the wholesale rejection of Genetically Modified Crop technology by the British public was significantly
at odds with the views of many scientists who chose to see the technology as safe and controllable (Frewer 2003, Frewer et
al. 2004, Tait 2001a). Additionally, of course, the examination of motives within science can itself be questioned as the
ongoing pressures of funding, essential to the continuation of particular research lines, and the requirement for novelty in
research in general, essential to publication and career development in science, means that there is a strong science agenda
which may be at considerable variance with wider societal wants and needs.
This report describes a study in which the ecological literature was reviewed to identify evidence for the effects of
pesticides in natural environments. The views of specific stakeholders were then elicited in focus groups on the acceptability
of pesticide effects from normal agricultural use on crops, and a public opinion survey of more than 2000 respondents across
the UK was run to determine what the public value and their attitudes towards the use of pesticides. Finally, information
from these activities was combined to develop a coherent decision framework for pesticide authorisation that consistently
incorporates public values, when appropriate.
2. AIMS AND OBJECTIVES
The overall aim of this project was to establish the ecological and socioeconomic bases for defining Ecologically
Acceptable Concentrations (EACs). The specific objectives were:
1. To produce a comprehensive literature review on current knowledge and understanding of the ecological
significance of long-term risks from pesticides to non-target species.
2. To run focus groups with particular stakeholders and to perform a statistically robust public opinion survey to
determine the spectrum of views on EACs and the factors that influence these views.
3. To combine information from 1 and 2 above within a formal decision analysis framework to provide PSD with a
coherent, logical, yet flexible decision tool for identifying EACs under Directive 91/414/EEC.
SID 5 (2/05)
Page 4 of 30
3. REVIEW OF SIGNIFICANCE OF LONG-TERM ECOLOGICAL RISKS FROM PESTICIDES TO
NON-TARGET ORGANISMS
3.1
Long term implications of pesticide exposure for ecosystems
3.1.1
Habitat and resource modification
Can plausible inferences be made about long-term habitat or resource modifications caused by the approved use of
pesticides, based on evidence in the literature? Without doubt, agricultural activities have produced long-lasting changes on
the British landscape and associated biota. The specific suggestion that pesticides play a significant role in habitat and
ecological resource change is supported by the following theory and evidence.
1.
The community conditioning hypothesis (Matthews et al. 1996) suggests that community structure may remain in an
altered state for a long time after pesticide residues have dropped to toxicologically insignificant levels. Sheffield
and Lochmiller (2001) provide an example of modified species interactions that persisted for longer than pesticide
residues.
2.
The removal of a keystone or dominant species alters habitat qualities (Power et al. 1996). Although it is important
to prevent this (Mills et al. 1993), current toxicity tests performed during the authorisation of pesticides cannot
ensure that keystone species populations remain extant. Therefore habitat or resource alteration due to unintentional
removal of keystone species could occur.
3.
Although commonly invoked, the general application of the ecosystem redundancy hypothesis has not been
adequately addressed and an equally plausible hypothesis, the rivet popper hypothesis, according to which the loss
of any species is of concern, might be a more accurate and conservative depiction of ecosystem consequences of
pesticide exposures (Pratt and Cairns 1996).
4.
British studies of grey partridge (Perdix perdix) provide strong evidence that herbicide use has diminished habitat
quality, and consequently, decreased partridge population densities (Blus and Henny 1997, Ewald and Aebischer
1999).
5.
Population densities of thirteen farmland bird species have fallen 30% on average in the United Kingdom from 1968
to 1995 but twenty-nine generalist species have increased (Krebs et al. 1999). This suggests a decrease in habitat or
resource quality for bird species that prefer farmlands.
6.
Other instances of medium-term adverse effects of agricultural practice on aquatic species (e.g., Furse et al. 1995)
and terrestrial species (e.g., Martin et al. 2000) have been reported, suggesting that longer-term effects are plausible.
In contrast to the above, the following theory and evidence detracts from any definitive statement that pesticides play an
important role in general changes to the farming landscape and associated biota.
1.
2.
3.
3.1.2
Most communities display a degree of resistance to change and an ability to recover after stress-induced change
(Cairns and Niederlehner 1993, Pratt and Cairns 1996).
Although current toxicity tests in support of pesticide authorisations generate imprecise predictions of population or
community vitality or viability, there are indications that effects information from these tests are, on average,
conservative relative to population effects (e.g., Caslin and Wolfe 1999, Bishop et al. 2000, Forbes and Calow
2003).
Other plausible causal agents co-occur with pesticides, such as physical habitat modification, introduced species,
and increased fertilizer use. The presence of these other potential causes confounds the identification of pesticides as
a major factor contributing to long-term habitat or resource modifications.
Local extinction and reduced biodiversity
Evidence can also be compiled from the literature about the risk of local extinction and reduced biodiversity. Support for
the view that pesticides contribute to local extinctions or reduced biodiversity includes the following:
1.
Laboratory studies have identified mechanisms capable of disrupting species interactions and, consequently, species
persistence and community structure.
2.
Publications addressing widespread declines in bird (Beaumont 1997, Krebs et al. 1999) and amphibian (Wake
1991) species suggest that pesticides might play a role. Sparling et al. (2001) provide one example of amphibian
decline that is directly linked to pesticide use.
3.
Mineau (2002) suggests that bird mortality likely occurs frequently in Canadian agricultural fields.
4.
Pesticides can modify endocrine function in ways that impact population viability (Gray et al. 1998). For example,
changes in alligator sexual characteristics and reproductive viability in Florida lakes are notionally linked to
endocrine-modifying pesticides (Crain et al. 1998, Guillette et al. 1996, 1999, Milnes et al. 2002).
5.
Pesticides can adversely influence vital rates, and changes in these vary in complex ways among species (Forbes
and Calow 2003).
SID 5 (2/05)
Page 5 of 30
6.
Many indirect effects, not quantified in conventional tests, are documented for diverse taxa exposed to pesticides.
These effects could lead to local extinctions and, gradually, to reduced diversity (Preston 2002).
7.
Although commonly invoked, the general applicability of the ecosystem redundancy hypothesis has not been
adequately ascertained. The equally plausible, and more conservative, rivet popper hypothesis might provide a more
appropriate framework for protecting ecosystems (Pratt and Cairns 1996).
8.
Long-term risk to mammalian species has been documented (Muller et al. 1981).
9.
Some mesocosm studies (Van den Brink 1996, Woin 1998) document species abundance shifts as a consequence of
pesticide exposure.
10. Bird populations are impacted by both legacy pesticides (e.g., Hickey and Anderson 1968) and modern pesticides
(e.g., Schroeder and Sturges 1975, Savidge 1978, Blus and Henny 1997, Martin et al. 2000, Mineau 2002, Sotherton
and Holland 2003). Some impacts have long-term consequences for bird populations.
In contrast to the above, theory and evidence detracting from any conclusion that pesticides influence local extinction
and reduced biodiversity are the three provided above for habitat and resource modification, and the following,
1.
2.
3.
3.1.3
Phenotypic plasticity and life history strategies are important features of most populations (Stearns 1992) and could
lessen the effects of pesticides on populations (Stearns and Crandall 1984, Sibly and Calow 1989).
Harrington and MacDonald (2002) saw no indication in the sparse data available that the decline in UK mammalian
species is linked to pesticides.
Some bird studies suggest that sublethal effects on nestling (e.g., Bishop et al. 2000) and adult birds (Blus and
Henny 1997) occur, but compensation results in these effects having no apparent adverse consequences on overall
fitness.
Recovery and recolonisation
Evidence supporting a negative view about community recovery and recolonisation after pesticide exposure includes the
following.
1.
2.
3.
4.
There can be a significant delay between the removal of a keystone species and its replacement in a community
(Ernst and Brown 2001).
Negative impacts on mammalian populations can persist longer than pesticide residues do in treated environments
(Sheffield and Lochmiller 2001).
Although life history traits can help identify mammal species that rely on dispersal to remain viable (Fagan et al.
2001), such traits are not formally considered in present risk assessment methods.
The community conditioning hypothesis (Matthews et al. 1996) suggests that community structure can remain
altered after pesticide residues have dropped to toxicologically insignificant levels.
In contrast to the above, evidence detracting from any conclusion about the adverse effects of pesticides on community
recovery and recolonisation include all of those provided above for habitat and resource modification, and local extinction
and reduced biodiversity.
3.2
An ecological vantage on pesticide risk assessment
3.2.1
Organismal
Most of the lower tier information applied in pesticide risk assessment is interpreted from an autecological vantage.
Such a vantage is directly useful for endangered or threatened species because the taking of even one of these individuals is
illegal, so information about direct effects on individuals is essential. Such information might also be relevant for
charismatic species for which the taking of an individual would be very undesirable. For example, a 1984 poisoning of
Atlantic brant (Branta bernicla) led to the US regulatory authorities discontinuing diazinon use on turf grasses despite the
lack of any evidence of risk to populations of this unlisted species (Bascietto 1998). Application of diazinon to golf courses
and sod farms was regarded as constituting an unreasonable risk to individuals of a charismatic species.
The autecological vantage is also useful, but insufficient, for predicting effects on species populations. This
approach has a long standing in ecology where it was used to examine the relationship between individual organisms or
species and their physical, chemical, and biological environment. Liebig’s law of the minimum (Liebig 1840), Shelford’s law
of tolerances (Shelford 1911, 1913), and the concept of a fundamental niche grow out of this premise that knowledge of the
tolerances or requirements of individuals can be used to predict species distributions and abundances. However, this vantage
is insufficient for explaining or predicting all aspects of population or community ecology, and the synecological vantage is
an essential one for general prediction or description of ecological systems (Preston 2002). This view can perhaps be
summed up by a quote from over twenty years ago: Although this discussion may appear hostile to single species toxicity
testing efforts, it is not intended to be. Single species tests are exceedingly useful and are presently the major and only
reliable means of estimating probable damage from anthropogenic stress. Furthermore, a substantial majority, perhaps
everyone in this meeting is certainly aware of the need for community and system level toxicity testing. How then does one
account for the difference between awareness and performance? (Cairns 1984)
SID 5 (2/05)
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3.2.2
Population
The EC Guidance Document on Aquatic Ecotoxicology (EC 2002a) and EC Guidance Document on Terrestrial
Ecotoxicology (EC 2002b) explicitly state that species populations are to be protected by the EC risk assessment process for
pesticides. Although the issue is gradually being remedied, the tools most often used in pesticide risk assessment provide
imperfect insights into population level effects. For example, Grant (1998) indicates that substantial changes in some vital
rates such as those often measured in toxicity tests can have little impact on populations because density-dependent
population dynamics shift to compensate for the toxicant’s effects.
Newman (2001) identified four reasons why current toxicity tests might not provide sufficiently accurate predictions
of consequences for field populations. First, toxicity testing often focuses on the most sensitive stage in an organism’s life
cycle based on the assumption that protection of individuals in this most sensitive stage will ensure protection of the
associated species population. But the most sensitive stage of a life history may not be the most important one relative to
maintaining a viable population, as illustrated by Kammenga et al. (1996). Second, metrics such as the LC50, LOEC or
NOEC cannot be incorporated directly into demographic models used to project population change through time. Third, postexposure mortality is ignored in most laboratory-derived metrics but field populations can experience significant mortality
after exposure to a toxicant ends (e.g., Newman and McCloskey 2000, Zhao and Newman 2004). Fourth and finally, the
underlying assumptions for some conventional concentration-effect models have not been resolved. As an example, the
probit model is based on the concept of the individual effective dose or concentration. It has been hypothesized that each
individual has a unique concentration above which it will die and below which it will survive exposure (Bliss and Cattell
1943, Finney 1947). The distribution of tolerances in any population is assumed to be lognormal. An alternate hypothesis is
that each individual has the same chance of dying as any other and whether it dies depends on chance alone. Which of the
hypotheses dominates in a particular population exposure scenario currently remains ambiguous. The consequences of this
ambiguity are significant because the two hypotheses predict very different population consequences with repeated exposures
(Newman and McCloskey 2000).
A risk assessment for atrazine by Solomon et al. (1996) provides another illustration of subtle but crucial
population-related issues left unexplored in current ecological risk assessment (ERA) approaches. Although not considered
in the very thorough ERA by Solomon et al. (1996), Hayes et al. (2002a,b, 2003) and Withgott (2002) indicate that atrazine
acts as an endocrine modifier and, more controversially, is suspected to affect the reproductive fitness of amphibians at
environmentally realistic concentrations. Dodson et al. (1999) found that atrazine also influenced the production of males in
Daphnia pulicaria cultures, and that D. pulicaria fecundity and survival were much less sensitive than male production.
Both of these effects could significantly change population vitality but were not considered in the original ERA for atrazine.
Solomon et al. (1996) focused on the effects of atrazine on primary production, maintenance of macrophyte community
structure, and long-term viability of fish populations in their higher tier ERA. In their effects characterisation, they
commented on the lack of information for amphibians but concluded that, “the limited data suggest that amphibians are
tolerant of atrazine.”
Appropriately designed laboratory (e.g., Van Der Hoeven and Gerritsen 1997, Snell and Serra 2000), mesocosm
(e.g., Van Den Brink 1996, Sherratt et al. 1999), enclosure (e.g., Caslin and Wolfe 1999, Wang et al. 2001), and field (e.g.,
Schroeder and Sturges 1975, Savidge 1978) studies can provide valuable insights into the population effects of pesticides,
and more studies of these types are included in both predictive and retrodictive pesticide risk assessments each year. As
more such tests are applied to ERA, the strength of associated inferences about long-term effects of pesticides to populations
will improve. As is occurring now in conservation biology, more and more emphasis is slowly being placed in ERA on the
risk of local extinction under specified exposure conditions (Tanaka 2003).
3.2.3
Community and species assemblage
As discussed already, interactions among species populations are also essential in any ERA to understanding the
long-term consequences of pesticide use. The value of individual-based effect metrics from the laboratory is ambiguous for
this purpose and, in some cases, demonstrably inadequate. Laboratory-based designs for quantifying community effects exist
(e.g., Cairns et al. 1986) but lack the realism of field or mesocosm studies. Field studies have tremendous value for this
purpose and methods exist for extracting insights about effects on species assemblages or communities (e.g., Savidge 1978).
Studies involving mesocosms or enclosures (e.g., Liber et al. 1992, Sheffield and Lochmiller 2001) may also provide
valuable information about these effects but the practical temporal scale is shorter for most mesocosm and enclosure studies
than for field studies so that, although such studies are more realistic than laboratory studies, they still may not reflect the
true field situation with adequate accuracy or precision (Crane 1997).
3.3
A conservation biology vantage on pesticide risk assessment
3.3.1
Pesticides and non-target species of conservation concern
Changes in the way that farmland is managed have unquestionably reduced insect abundance over the last 30 years
in the UK and other countries (Wilson et al. 1999, Benton et al. 2002). Pesticides could have contributed to this decline both
through direct poisoning by insecticides (Dover et al. 1990; Cardwell et al. 1994) and the removal of food plants by herbicide
application (Marshall et al. 2003). Many bird species declining on farmland are reliant on insects and other invertebrates
during the breeding season as staple nestling food (Cramp 1988, 1992, Cramp and Perrins 1993, 1994a, 1994b, Benton et al.
2002, Gruar et al. 2003). Furthermore, there is a correlation between the declining abundance of insects on farmland and the
abundance of several bird species (Benton et al. 2002). Such a correlation, however, is not definitive proof that pesticide
SID 5 (2/05)
Page 7 of 30
usage is the principal cause of the decline. There is ample evidence that agricultural intensification is responsible for
reducing bird populations on farmland (Chamberlain et al. 1999; Henderson et al. 2000) and, while pesticide usage has
evolved during the intensification process, it is not the only practice that has changed. For example, the UK corncrake (Crex
crex) population crashed during the 20th century through the intensification of farmland practices (Green and Gibbons 2000),
but here the driving factor was the shift from hay cutting to silage production (Green and Stowe 1993, Stowe et al. 1993).
Silage is cut much earlier in the year than hay with the consequence that corncrakes, which nest in long grass, are unable to
produce fledglings before the nests are destroyed during mechanical cutting of grass (Broyer 1994, Green 1996). The skylark,
Alauda arvensis, and lapwing, Vanellus vanellus, provide further examples of species that have declined on farmland through
reasons other than pesticide application (Chamberlain et al. 2000, Wilson et al. 2001).
There are few examples where pesticides are known to have affected the population sizes of non-target species in
the long term. Exceptions include the bioaccumulation of organochlorine (OC) pesticides through food chains into top avian
predators, such as the peregrine falcon, Falco peregrinus (Ratcliffe 1980) and the sparrowhawk, Accipiter nisus (Newton
1986), and the effect of the loss of insects in field margins on the grey partridge (Chiverton 1999), as described earlier. Once
the effect of OC pesticides was realized, appropriate legislation was introduced in developed countries and the bird of prey
populations recovered to earlier levels (Millsap et al. 1998, Wingfield Gibbons et al. 1993, Horne and Fielding 2002). For the
grey partridge, the introduction of conservation headlands conclusively demonstrated a link between pesticide application
(both herbicide and insecticide) and a reduction in insect food required by the precocious grey partridge chicks (Green 1984,
Borg and Toft 2000, Southwood and Cross 2002). The headland experiments further showed that pesticide application could
be sympathetically designed to achieve both acceptable agricultural and conservation targets. Where implemented, headlands
have resulted in a dramatic increase in the local abundance of grey partridges (Rands 1985, 1986; Chiverton 1999). In
addition, a remarkable increase has also been recorded in the numbers of certain very rare species of farmland flowers, such
as pheasant’s-eye, Adonis annua, shepherds needle, Scandix pectin-veneris, and cornflower, Centaurea cyanus. Headlands
are also more heavily used by butterflies (Dover et al. 1990) and small mammals (Tew et al. 1992).
Despite a general lack of direct evidence, pesticides and veterinary medicines are frequently cited as possible factors
causing population declines in currently scarce species. For example, the hornet robberfly, Asilus crabroniformis, breeds on
farmland and is associated with large grazing animals (Holloway et al. 2003a). In Britain, the species is Biodiversity Action
Plan listed and scarce, and exists as a series of fragmented populations (Smith 2000, Clements and Skidmore 2002).
Ivermectin, a veterinary medicine, is cited as a factor contributing to the decline of hornet robberflies (Smith 2000). While
research has shown that ivermectin can influence the numbers of coprophagous flies emerging from dung (McCracken and
Foster 1993, Wardhaugh et al. 2001), the effect is considerably less dramatic for dung beetles (Kadiri et al.1999). By
extension, it is therefore assumed that ivermectin can negatively influence the hornet robberfly (Clements and Skidmore
1998, Smith 2000, 2001), even though neither a direct nor indirect effect has ever been shown. There is a general assumption
within the conservation community that pesticides and other anthropogenic chemicals are detrimental to wildlife. Concern
about pesticide application extends mainly to the effects of spray drift. For example, where pesticide application is occurring
close to a Site of Special Scientific Interest, a barrier such as scrub, is often allowed to develop between the two sites to
intercept any drift (D. Sheppard, English Nature, pers. comm.) as a precaution against any possible adverse effects.
Many species have become rare on farmland and beyond as a result of habitat loss and degradation. Consequently,
many populations are highly fragmented across the landscape substantially altering plant (Jacquemyn et al. 2003, Luoto et al.
2003) and animal (Sherman and Runge 2002, Schmiegelow and Monkkonen 2002, Ryttman 2003, Bellamy et al. 2003,
Watson et al 2003) population dynamics. Some populations, particularly of insect species such as silver spotted skipper,
Hesperia comma (Hill et al. 1999), Glanville fritillary, Melitaea cinxia (Saccheri et al. 1998) and bog fritillary, Proclossiana
eunomia (Sawchik et al. 2002), but also some vertebrate species such as red squirrels, Sciurus vulgaris (van Apeldoorm et al.
1994) and nuthatches, Sitta europaea (van Langevelde 2000), function as metapopulations, i.e., a series of populations only
loosely connected through occasional dispersion of individuals among populations (Hanski 1999). A situation that is also
common in the wild is where local populations have gone extinct and the distance separating populations is too great to allow
exchange of individuals through dispersion (Hill et al. 1996, Epperson 2000, Kudoh 2001, Doebeli and Killinback 2003).
When this happens, genetic variation in an isolated population begins to erode at a rate negatively correlated with the size of
the population (Frankham et al. 2002). Small populations go through population bottlenecks and, particularly if the
population remains small for several generations (inevitable if a species is rare), a substantial amount of genetic variation can
be lost (Frankham et al. 2002). K-strategy species may be at more risk from extended bottlenecks than r-strategists due to
their slower rates of recovery (Begon et al. 1986). Notionally, lower genetic variation results in a decline in adaptability, an
increased probability that an entire population will be lost as a result of a single event, and a general reduction in fitness
(Ralls and Ballou 1983, Frankham et al. 2002). There are examples of species that survive in the wild with apparently very
little genetic variation, for example the northern elephant seal, Mirounga angustirostrus (Hoelzel et al. 1993, Hoelzel 1999)
and the fallow deer, Dama dama (Pemberton and Smith 1985) in Britain but, of course, those populations that have
disappeared as a result of reduced genetic variation no longer exist to allow their levels of variation to be examined. Madsen
et al (1999) working with adders, Vipera berus, unequivocally demonstrated that low genetic variation is associated with low
fitness for this species. With rare, fragmented populations in agricultural landscapes, pesticides could play a role in the longterm reduction in fitness of some populations by repeatedly driving them through bottlenecks. Each time this happens, the
genetic variation would be reduced more and more until the population enters an extinction vortex (Tanaka 2000, Lane and
Alonso 2001, Rowe and Beebee 2003, McGinnity et al. 2003) from which there is no prospect of recovery. In addition to this
process, if pesticide application results in the loss of one or more populations within a metapopulation, increasing the
distance among populations to beyond the maximum for successful emigration, populations could become isolated and enter
a process of genetic erosion.
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3.3.2
Future considerations
The risk of pesticide application to, for example, vertebrate wildlife is currently assessed using small numbers of
bird (usually Japanese quail, Coturnix coturnix, or Northern bobwhite quail, Colinus virginianus, and mallard ,Anas
platyrhynchos) or mammal (e.g., rat, Rattus norvegicus or mouse, Mus spp.) species under laboratory conditions. It is fair to
say that pesticide risk assessment requires a leap of faith to extend the results from the small number of species tested under
artificial conditions to a natural and complex field scenario. Furthermore, proposals for the safe use of pesticides under field
conditions often focus on species of economic value. For example, to protect honeybees, Apis melifera, it is recommended
that spraying be carried out early in the morning or late in the evening to coincide with the cooler temperatures before and
after flight. However, bumblebees Bombus sp. and many species of solitary bees are able to fly under cooler conditions than
honeybees and become non-target casualties of pesticide application. It is difficult to integrate pesticide considerations into
conservation because of the multitude of possible effects. Consequently, there is a general precautionary assumption made
that pesticides applied for agricultural purposes are detrimental to wildlife and that sites valuable to conservation, for
whatever reason, should be protected from pesticides [although note that pesticides are sometimes applied for conservation
purposes, for example to control invasive weed species (http://www.english-nature.org.uk/pubs/Handbooks/default.asp)]
It is possible that the role that pesticides might play in the reduction of animal and plant populations, through
whatever routes, may come under closer scrutiny in the future for several reasons. First, the increased awareness of long-term
genetic consequences of short-term poisoning events may focus more attention on the distribution of pesticides across
farmland habitats. Another issue is global warming (Hulme and Viner 1998) resulting in weather pattern changes in many
countries. It is quite possible that the growth rate of arable weeds could increase considerably in temperate regions as a result
of warmer, and probably wetter, conditions at particular times of the year (Hossel 2001). Pest insects might also survive in
higher numbers. For example, the winter of 1989/90 was exceptionally mild in the UK, which triggered a big increase in the
number of aphids spreading disease to winter cereal crops. Crop yields fell by as much as one tonne per hectare (Hossel
2001). Global warming is likely to result in increased pesticide application. However, the general public have become more
aware of the amount of toxic substances used in food production and an increase in the quantities applied is likely to meet
opposition, so outcomes are difficult to predict.
Another factor that may result in increased scrutiny of pesticides is governmental activity such as the biodiversity
strategy for England entitled ‘Working with the Grain of Nature’ (http://www.defra.gov.uk/wildlifecountryside/ewd/biostrat/). The target is to ensure that species are part of ‘healthy functioning ecosystems’ and to ensure that
‘biodiversity considerations become embedded in all the main sectors of economic activity.’ With a focus on ecosystems
comes an acceptance that species exist within complex food webs and communities (Giler and O’Donovan 2002). A change
of practice affecting one component of such a community could have a cascade effect all the way up to a species of
conservation concern. Agricultural practices are considered to be very important within the context of UK government’s
overall vision for the country’s biodiversity.
3.3.3
Long-term effect data required by conservation community
There is a paucity of data that unequivocally demonstrates a link between pesticide use and long-term population
reductions in farmland environments. For example, most of the species of conservation concern in the UK are not associated
with farmland (www.ukbap.org.uk) and it is generally accepted that, for the majority of species, habitat loss or alteration is
clearly the most significant factor threatening species extinction at the moment. Consequently, effects of pesticides are rarely
considered. Part of the reason for this is that following a poisoning event, long-term monitoring is infrequently carried out
and the potential ways in which pesticides could alter community structure are not studied. The investigation of both of these
issues is limited by the amount of available resource and with so many immediate problems, long-term or more complex
considerations fall low on the agenda. Furthermore, with the political desire in many countries to see wildlife conservation
largely directed by local communities, the scientific expertise to appreciate and investigate the more insidious, but
nevertheless potentially damaging, changes is not usually available. This type of research is the realm of a few academics
within universities or overstretched non-government organizations.
Management of farmland to consider biodiversity is generally voluntary and often entails joining an incentive
scheme, such as the Countryside Stewardship Scheme in the UK (http://www.defra.gov.uk/erdp/schemes/css/default.htm).
The way that farmland within the scheme is managed is dependent on the conservation objectives of that part of the
countryside. Within such schemes, particular pesticide application regimes need to be approved and here knowledge of the
longer-term effects of pesticide application would be very useful. For example, spraying with a given active substance may
result in the decline of a suite of butterfly species, but how long would it take for the butterfly community to recover? Armed
with this information, regulatory authorities may be able to recommend a suitable period of time between subsequent sprays
to maintain both the butterflies and to achieve the pest control objective. Alternatively, a pesticide may be particularly
detrimental to a small number of species that could affect community structure and ultimately a species of conservation
concern. For example, a proposed herbicide application could eliminate food for meadow grasshoppers, Chorthippus
parallelus, in field margins in an area. This species is not rare, but its powers of dispersal are very limited (Chinery 1993) so
it could take some years for the area to be recolonized. Meadow grasshoppers are the staple food source of growing cirl
bunting (Emberiza cirlus) nestlings, so the proposed application regime could ultimately eradicate a rare, target species
through an indirect route.
In conclusion, there are very few examples known to conservationists in which pesticides have been shown to have
direct or long-term damaging effects on wildlife. However, potential effects of pesticides are rarely considered in
conservation and pesticide issues, for example in connection with nature reserves, are examined on a case-by-case basis. The
conservation community would welcome a general reduction in the use of pesticides on farmland, often as a result of the
perceived impact of their use rather than direct evidence. Reductions are most likely to be achieved through government or
SID 5 (2/05)
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consumer pressure, e.g., the move towards organically grown foods (de Boer 2003). An integration of pesticide application
and effects into conservation strategy would be desirable, but can only be realized if appropriate data are collected. Currently,
these data are generally unavailable.
3.4
Quantitative tools for population and community analysis
3.4.1
Population analyses
Population effects can be quantified in a variety of ways. Application of demographic methods to chronic Daphnia
test data is possible with only minimal changes to current methods (Newman and McCloskey 2002, Section 4.3 of
ECOFRAM 1999). Nacci et al. (2002) applied a demographic approach to analyzing toxicant exposed field populations.
Caswell (1996) describes a straightforward approach to performing demographic analyses for toxicant-exposed populations
and, in his book (Caswell 2001), provides many details about the matrix approach to demographic analysis. ECOFRAM
(1999, Section 4.4 of the ECOFRAM Aquatic Report) provides general details for implementing population models in
studies of population effects. With knowledge of the relevant exposure duration, a stochastic projection of population
dynamics over the exposure period can be used to estimate the risk of local extinction at different exposure concentrations,
i.e., a statement of population risk can be made with these methods (Newman 2001). If spatially-explicit modelling is
required, such an approach can be taken with a metapopulation model (O’Connor 1996, Newman 2001). Mackay et al.
(2002) provide an example of a spatially explicit, population-based ERA.
Several software packages are available for analysis of demographic information. Population Viability Analysis
(PVA) has been available as a tool in wildlife conservation for over a decade (Soulé 1987), and involves the estimation of
extinction probabilities through analyses that incorporate identifiable threats to population survival into models of the
extinction process. Models, such as VORTEX, are available to carry out PVA (Lacy 1993) and there are many examples
where PVA scenarios have been modelled and tested (e.g., Ball et al. 2003, Kaye and Pyke 2003). Studies have also been
carried out using PVA to provide recommendations for the survival of rare vertebrates, such as the eastern barred bandicoot
Perameles gunnii (Lacy and Clark 1990), and invertebrates, such as Fender’s blue butterfly, Icaicia icaroides fenderi
(Schultz and Hammond 2003). While PVA has undoubtedly great potential use in conservation, it has yet to be widely
adopted as a tool by practitioners. The RAMAS program (Ferson and Akhakaya 1990) is another inexpensive program useful
for deterministic and stochastic modelling of populations and PVA, and is available in a version which explicitly considers
ecotoxicological information.
A further shareware example of a population modelling tool is the PopTools (http://www.cse.csiro.au/poptools) addin to ExcelTM which implements demographic methods to fit population data, and to make deterministic or stochastic
population projections.
3.4.2
Community and assemblage analyses
The conventional approach to effects characterization produces a series of relevant effect metrics such as LC50
values for test species for comparison with expected exposure concentrations. Recently, a probabilistic approach has emerged
for many risk characterizations beyond Tier I.
There are considerable differences in sensitivities among species and a focus on the most sensitive of a small
number of tested species does not take full advantage of the available data. Species sensitivity distributions (SSDs) make
fuller use of the available effects data. An SSD is a distribution of effect metrics for individual species thought to represent
collectively the species of concern. Effect metrics for the test species are ranked from lowest to highest, and their ranks
converted to approximate proportions. The paired proportions and effect metric concentrations are then fitted to one of
several models (Posthuma et al. 2002).1 Some SSD models include all species; however, some require separate models for
taxonomic subsets of species. For example, Solomon et al. (1996) performed a retrodictive ERA for atrazine in the North
American cornbelt by exploring SSD models for logical species groupings (e.g., Figure 21 in Solomon et al. (1996)).
In many recent “probabilistic” risk assessments, the creation of a SSD relies heavily on one of several statistical
distributions such as the log logistic, lognormal, or triangular distribution. For aquatic risk assessments of pesticides, data
points in these distributions are taken from endpoints in acute or chronic toxicity tests. In acute distributions, data points are
normally taken from tests used to derive LC/EC50 values. For chronic toxicity distributions, no-observed-effectconcentrations (NOEC values) are commonly used. In acute toxicity tests, exposures to a contaminant are generally of short
duration (e.g., 24 to 96 hours) and chronic toxicity tests are conducted over a full life cycle or an early life stage of an
organism.
For a pesticide ERA, it is important to identify a threshold hazard concentration above which ecological effects are
likely to occur. With SSDs, this is often approximated by selecting a low centile of the distribution. The resulting metric is
commonly called a hazardous concentration (HCp). Normally, the fifth or tenth centiles (HC5 and HC10) have been arbitrarily
used in ERA (Aldenberg and Slob 1993, Wagner and Løkke 1991). These lower centiles of an effect concentration
distribution have been applied historically in deriving US water quality standards and, in the case of the HC5, have more
recently been recommended in the EU Technical Guidance Document. Software available for fitting SSDs and estimating
HC5 values has been developed by Van Vlaardingen et al. (2003).
As discussed by the Aquatic Dialogue Group (1994), a risk assessment that relies solely on the protection of a
certain proportion of exposed species might not be protective if keystone, dominant, or legally protected species are ranked
below the specified proportion on a SSD. In choosing a proportion from a distribution of acute or chronic effects, one makes
the assumption that “protecting” a certain proportion of species will be protective of the structure and function of an
1
Grist et al. (2002) describe a nonparametric method that circumvents the need to identify a well-fitting distributional model.
SID 5 (2/05)
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ecosystem, and that the available single-species toxicity tests are representative of the ecosystem to be protected or the
universe of species in the environment. In reality, it would be remarkably fortuitous if the issue of protecting crucial species
were adequately addressed in laboratory testing of small numbers of conventional species. Also, the argument could readily
be made that acute LC50 or chronic NOEC information is not adequate for predicting a concentration to protect a species
population existing in a natural community (Hopkin 1993, Jagoe and Newman 1997, Newman et al. 2002, Newman and
Unger 2003). Regardless, it is now a common risk assessment practice and Maund et al. (2001) introduced some supporting
evidence for using the tenth centile of acute distributions based upon ecologically significant effects observed at higher
concentrations in field studies.
Effect quantification for population and species assemblages are less common than those for individuals yet, based
on the materials discussed to this point, the need for such metrics is high in pesticide risk assessment. Fortunately, relevant
methods are being applied more and more frequently, and higher tier methods such as mesocosm experiments, enclosure
studies, and field surveys are amenable to their use.
Community effects can also be extracted from laboratory, enclosure, mesocosm and field studies using conventional
ecological methods. Newman (1995), Matthews et al. (1998), and Clements and Newman (2002) provide information
specific to their application for the risk assessment of chemicals such as pesticides. The influence of simple community
interactions such as predator-prey interactions can be quantified (e.g., Tagatz 19762) in laboratory assays but this is not often
done to support ERA activities. More commonly, mesocosm community structure metrics are used in predictive pesticide
risk assessment activities and field community structure metrics are used in retrodictive risk assessment activities. The most
common indices are species richness, diversity and equitability indices. Communities or species assemblages might be
compared for different exposures using distance metrics. Multivariate methods such as ordination or clustering methods can
assess differences and similarities in species assemblages that have different exposure histories. All of these methods rely on
community structure information such as species abundances or presence/absence data.
Multimetric methods can incorporate structural and functional qualities of species assemblages during assessments
of effect (Clarke 1999, Clements and Newman 2002). The most common multimetric index is the index of biological
integrity (IBI) (Karr et al. 1986, Karr 1991, 1993). Karr’s IBI attempts to quantify the integrity of a system of concern
relative to an undisturbed or intact system of the same type in the same geographical region. As such, the IBI score for a
system has quantitative meaning only in comparison to that of an undisturbed system. The IBI concept has been applied
successfully to summarize the integrity of numerous sets of mesocosm or field data.
Numerous software packages implement conventional ecological metrics and multivariate methods. One example of
such shareware is the BioDiversity package available from Neil McAleece (biodiversity@nhm.ac.uk) of The Natural History
Museum and Scottish Association of Marine Science. Other frequently used software for multivariate analysis of ecological
community data includes PRIMER (Plymouth Routines In Multivariate Research; see http://www.primer-e.com/in) and
CANOCO (see: http://www.plant.dlo.nl/default.asp?section=products&page=/products/canoco/right.htm). Pastorok et al.
(2002) provide a comprehensive review of available modelling tools and software for analysis of chemical effects on
populations, ecosystems and landscapes.
A development in wildlife conservation that followed the Bern Convention (1979) was the realization that, in order
to conserve species, there is a need to address the state of the habitats in which they exist (http://www.englishnature.org.uk/baps/habitats/), because the two are inextricably linked. Attention to habitat and landscape issues has provided
a platform for the use of Geographic Information System (GIS) software in wildlife conservation. Geographic Information
Systems have been used in geographical studies for some time, but possible applications in risk assessment have only
recently been appreciated. As with PVA, there are examples of potential uses of GIS in conservation and pesticide risk
assessment (e.g. Markus et al. 2003, Holloway et al. 2003b), but its application has yet to be fully embraced by practitioners.
3.4
Discussion and conclusions
It is difficult to envisage a risk assessment process capable of consistently predicting long-term pesticide effects
founded primarily on individual-based effect metrics. A risk assessment process that includes more population or
community effects metrics would substantially reduce uncertainty in predicting long-term ecological effects of pesticide use.
The best illustration of this point is the current assumption that conservative calculations using individual-based effect
metrics allow conservative expression of protection of communities from unacceptable risk. In many cases, a predictive risk
assessment progressing to higher tiers would not include any information on indirect effects that might be common and
important in ecological systems. Furthermore, laboratory test species tend to be r-strategy species. This condition creates
high uncertainty about predictions of field population persistence for K-strategy species.
Can the current ERA process form the foundation for estimating long-term ecological risks of pesticide use?
Current tests do provide useful information about the direct effects of pesticides. It would be unwise to abandon these tests
completely and to require only complex and expensive ecosystem studies during the predictive assessment of pesticides.
However, more laboratory tests focused on population level effect metrics can be performed and would generate more insight
than currently possible. For example, risk predictions from the species sensitivity distribution (SSD) approach would be
greatly improved by using risk of local population extinction instead of risk of exceeding an LC50 value. It is currently
impractical to attempt this with most published information from laboratory toxicity tests performed to date, either because
information on time-dependent survival and fecundity is not collected, or because it is not reported in sufficient detail.
2
Table 2.6 in Clements and Newman (2002) gives details for several predator-prey experiments conducted to quantify the
effects of toxicants on species interactions. Six involve pesticides as the stressor.
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Another improvement would be if strict adherence to the requirements of legislation such as Directive 91/414, with
listing of the results of sometimes inappropriate toxicity tests, were replaced by more formal adoption of the principles of
ecological risk assessment. This would include formal construction of a conceptual model for each active substance of
exposure sources and pathways, and identification of potentially sensitive receptors in Tier 1, rather than only at higher tiers,
if at all. In drafting the risk assessment, as much effort should be spent on integrating known ecological relationships into the
assessment as is presently being spent on the compiling of concentration and individual-based effects data. A rich literature
on ecology remains grossly underexploited in pesticide risk assessment activities.
The continuing use of mesocosm or enclosure studies might also result in more ecologically relevant information for
conducting predictive risk assessments because of an increased likelihood of seeing an indirect effect of a pesticide before it
is authorized and enters into wide use. However, most such studies currently performed in Europe omit vertebrate species
because their extensive feeding can confound measurements of invertebrate and plant populations in mesocosms. This
means that indirect effects on vertebrates remain unmeasured, and those that are observed in mesocosms may be due to the
absence of top predators. The realism of mesocosms when predicting the effects of pesticides therefore remains uncertain.
Finally, and of great importance, more post-authorization monitoring could serve as a safety net for the predictive
risk assessment of pesticides, which cannot assess all plausible direct and indirect effects for all systems or pesticide
mixtures. Pesticides are currently authorized for use in most countries by combining toxicity data with modeled predictions
of exposure concentrations. There is some evidence that model scenarios combine conservative assumptions that do not
occur widely in the environment (Hendley et al. 2001). Kapustka et al. (1996) describe the 'ecological disconnect' that results
from the gulf between the current simple pesticide authorisation procedures and the complex ecological protection goals of
these procedures, and suggest that, 'It is remarkable that with such imperfect information, so few reports of pesticide
incidents occur. Alternatively, one could argue that the registration process is exceptionally conservative to the extent that
some beneficial uses are excluded without cause.' Their answer to the problem of ecological disconnect is to treat risk
estimates made during pesticide registration as working hypotheses. These would then require post-authorisation testing,
through field monitoring of concentrations and biological effects in the environment during a probationary use phase. This
seems to be a sensible and logical approach, and it is surprising that it has not been widely adopted by regulatory authorities.
Field studies to examine the biological effects of pesticides must be able to distinguish between effects caused by
natural stressors (e.g., extreme weather events), other anthropogenic stressors (e.g., habitat modification or exposure to nonpesticide chemical pollution), and natural covariates (e.g., the prevailing climate, geology and geography), as these may all
influence the structure of organism assemblages (Wickham et al. 1997). In addition to this, the effects of stressors can only
be determined in relation to a reference condition (Rykiel 1985), which is the condition that would occur in the absence of
stress.
Unfortunately, many reported field studies do not meet the underlying assumptions of hypothesis testing statistics
because they are unreplicated or pseudoreplicated, and unable to assign treatments at random (Beyers et al. 1995). For
example, comparison of upstream and downstream sites will likely include factors that covary with pesticide contamination,
such as other agricultural impacts (e.g., nutrients, sediments, or physical disturbance (Sallenave and Day 1991)).
Investigations performed in this way will almost inevitably attract criticism, especially when the politically charged subject
of pesticide use and effects is under study. To address such defects, Suter (1993) emphasizes the need for rigorous logic
when attempting to unravel the ecological epidemiology of pollution effects in field studies. Both he and researchers with
experience in field studies of pesticides (e.g., Beyers et al. 1995) identify Koch's postulates and Hill's factors, both
appropriated from medical epidemiology, as useful for focusing attention on the criteria needed to demonstrate causality in
field studies.
Koch's postulates, modified for use in environmental toxicology, are,
1. The effects of a toxicant must be regularly associated with exposure to the toxicant and any contributory causal
factors. According to Suter (1993), such a regular association should normally consist of Kant’s criteria for
causation (law of succession and concept of action). These are that cause and effect must always occur together, and
that the effect must follow, not precede, the cause.
2. Indicators of exposure to the toxicant must be found in the affected organisms. This could be established by either
measuring the toxicant in the organism or measuring a relevant biomarker induced by the toxicant.
3. The toxic effects must be observed when normal organisms or assemblages are exposed to the toxicant under
controlled conditions, and any contributory factors should contribute in the same way during the controlled
exposures. This criterion is best met through use of laboratory toxicity tests or mesocosms (e.g., Schulz et al. 2002)
to confirm that organisms are affected to the same degree at measured toxicant concentrations.
4. The same indicators of exposure and effects must be identified in the controlled exposures as in the field. Again, this
is best established through laboratory or mesocosm experiments to confirm that similar concentrations of toxicant
within organisms lead to similar effects.
Koch's modified postulates are augmented by Hill's criteria, which were used by Gilbertson (1997) and co-workers to
establish that organochlorine pollutants had adversely affected fish, other wildlife and humans in the Great Lakes basin. The
main elements of Hill's criteria are,
1. Specificity: Does only the potential cause lead to the effect, and does the potential cause lead only to the effect?
Meeting this criterion would be fortunate in most environmental investigations, as it is often the case that multiple
causes have multiple potential effects. However, a very high degree of cholinesterase inhibition, outside the normal
range of natural variability, would be an example of a specific effect that is very likely to be caused only by
exposure to organophosphorus or carbamate pesticides.
2. Strength of association: How precise is the relationship between the potential cause and the observed effect?
3. Time order: Does the effect follow the cause temporally? (Kant’s law of succession).
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4.
Consistency on replication: Is the association repeatedly observed at different times and places by different
investigators?
5. Coherence: Does a cause-effect interpretation of the data seriously conflict with generally known facts? Are there
plausible mechanisms of toxic action?
Design of studies around Hill's and Koch's criteria is a logical approach to answering questions about the link between
pesticides and effects in nature, as found by researchers studying organochlorine effects in the Great Lakes (Gilbertson
1997). We believe that this requires five main elements:
1. Studies should be extensive, rather than intensive, and include measurements at as many sites as possible, so that the
consistency of association between pesticides and effects can be assessed.
2. Pesticide concentrations in surface waters should be measured rather than inferred through modeling, so that the
presence of a causal agent can be demonstrated.
3. Ideally, researchers should possess detailed knowledge of pesticide application times, rates and locations near
particular aquatic sampling sites so that the time order of possible pesticide cause and effect can be examined more
accurately. This may not be necessary if organism exposure concentrations are well characterized in the field,
although it would almost certainly be of great use when identifying potential study sites.
4. Concentrations of pesticides in organisms should also be measured as a direct indicator of exposure.
5. The plausibility of potential pesticide cause and effect relationships in the field should be demonstrated by
laboratory studies to confirm that concentrations of pesticides found in the field can cause the observed effects on
particular taxa.
4.
ACCEPTABILITY OF PESTICIDE IMPACTS ON THE ENVIRONMENT: WHAT DO UNITED
KINGDOM STAKEHOLDERS AND THE PUBLIC VALUE?
4.1
Methods
Focus groups with specific stakeholders in the UK pesticides authorisation process were held to elicit a wide range
of views and questions about the acceptability of pesticide effects. The outcomes from these focus group discussions were
then used to formulate and prioritise questions for a public opinion survey.
4.1.1
Stakeholder focus groups
The following specific stakeholders were identified as key decision-makers or opinion-formers in the UK pesticides
authorisation process:
• The Pesticides Safety Directorate, who are the competent authority for Directive 91/414/EEC in the UK.
• Other government regulatory agencies with an interest in pesticides or nature conservation.
• Pesticides manufacturers.
• Environmental non-governmental organisations.
• Farmers and their advisors.
• Food distributors and retailers.
• Academics with knowledge of the fate and effects of pesticides.
• Environmental consultants with knowledge of the fate and effects of pesticides.
Representatives from each of these stakeholder groups (Table 1) were invited to separate 3-h meetings held during
December 2003 at which they were asked to discuss:
1.
In general, what levels of effect (lethal or sublethal) that may be caused by normal agricultural pesticide
use on crops are acceptable and what are unacceptable?
2.
Does your response depend upon the particular species affected?
3.
What environmental scale does your response relate to (e.g., a single field, a parish, a county, or the whole
country)?
4.
What frequency of occurrence does your response relate to (e.g., acceptable once a year, once every 10
years, etc.)?
The views expressed by the different stakeholder groups were collated for each group after the meetings and sent to
them for comment. Once group members had agreed on the summary, the eight individual summaries were consolidated into
one report which was then sent to all the stakeholders for any further comment, and final agreement.
Table 1 Stakeholder organisations represented at focus groups
Organisation Type
UK pesticide regulator
Pesticides manufacturers
Government
SID 5 (2/05)
Organisation Name
Pesticides Safety Directorate
Bayer Crop Science Ltd
Dow AgroSciences Ltd
BASF plc
Dupont (UK) Ltd
Syngenta Crop Protection UK Ltd
Environment Agency of England and Wales
Scottish Environmental Protection Agency
Central Science Laboratory
Veterinary Medicines Directorate
Broads Authority
Page 13 of 30
Organisation Type
Environmental Non-Governmental Organisation
Farmers & Advisors
Food Distributors
Academics
Consultancy
Organisation Name
English Nature
National Trust
The Allerton Research & Educational Trust
Pesticides Action Network-UK
Wildlife & Countryside Link
Farming and Wildlife Advisory Group
Random Pulse
Horticultural Development Council
Home Grown Cereals Authority
Farmer
Farmer
National Farmers Union
Association of Independent Crop Consultants
Velcourt Research & Development Ltd
Farm Retail Association
Fresh Produce Consortium (Farm Care)
University of Warwick
University of Sheffield
Centre for Ecology & Hydrology
Cranfield University
University College London
Ponds Conservation Trust
Cambridge Environmental Assessments
JSC International Ltd.
Braddan Scientific
Huntingdon Life Sciences
TSG Europe
Safepharm
Envirocorp
4.1.2
Public opinion survey
A representative quota sample of 2,049 adults aged 15+ in 201 sampling points across Great Britain was interviewed
by MORI. The interviews were conducted face-to-face, in-home between 29 April – 4 May 2004 and the data were weighted
to the known national population profile.
Questions for the public opinion survey were formulated and prioritised by stakeholders and are shown in Table 2.
Question 1 was asked separately, at the start of each interview, and Question 2 onwards was asked later in each interview.
Questions 3 and Questions 4/5 were alternated, as were Questions 6 and 7, to avoid bias. At Question 8, the sample was split
into two versions, with one half of the sample (Version One) asked about their concerns associated with the effects of human
activities on the countryside, and the other half (Version Two) asked about their concerns associated with the effects of
pesticides used on farm crops.
At Question 11 a simple choice experiment was designed to derive willingness-to-pay estimates for different
socioeconomic groups of the population. For this, respondents were each faced with one choice set consisting of two
hypothetical food baskets: one labelled ‘standard’ and one labelled ‘no pesticides’. Prices of the latter varied between £14,
£17 and £23 per week and were randomised across respondents, while the price for the former was fixed at £12 per week. As
a substitute reminder, each respondent was in addition provided with one of two weekly budgets (£30 or £50 per week) to be
allocated to both food and leisure. Choices in Question 11 were also alternated.
Table 2 Public opinion survey questions
Question
Q1. Which of these products do you use at home or in the garden?
Q2. Which of these are the two or three most important factors that
society should take into account in deciding how to grow food?
Q3. What possible causes of damage to wildlife and habitats in Britain
have you heard about?
SID 5 (2/05)
Choices
Rat/mouse poison
Weedkiller for paths, drive or patio
Weedkiller for lawn or other vegetated areas
Ant powder/spray
Wasp powder/spray
Slug/snail pellets
Sprays to protect plants from pests, fungus or disease
Organic pesticides
Other
None of these
Producing cheap food
Protecting animal welfare
Protecting jobs/Generating jobs
Protecting the beauty of the countryside
Protecting the health and variety of wildlife
Protecting waterways
Protecting human health
Producing good quality food
Other
None of these
Don’t know
Industrial chemicals/Use of industrial chemicals
GM crops
Housing/Building houses in the countryside
Page 14 of 30
Question
Q4. Do you personally support or oppose the use of pesticides on farm
crops?
Q5. Why do you say that? (asked only of those who supported or opposed
the use of pesticides in Q4)
Choices
Intensive farming
Litter/People dropping litter
Sewage discharges
Livestock/ The way animals are kept
Pesticides/Use of pesticides
Roads/Road building
Cars/traffic
Salmon farming/Fisheries
Other
None of these
Don’t know
Strongly support
Tend to support
Neither support nor oppose
Tend to oppose
Strongly oppose
Don’t know
Positive Codes
Kill unwanted pests
Control diseases of crops
Control plants that compete with
crops
Increase amount of food grown
Provide cheaper food
Q6. What are the good things about using pesticides on farm crops?
Q7. What are the bad things about using pesticides on farm crops?
Q8. How concerned are you personally about the effects of [human
activities in the countryside] or [pesticides used on farm crops] on each of
the following:
A. earthworms
B. butterflies
C. insects
D. songbirds (e.g. skylarks)
E. carrion birds (e.g. crows)
F. birds of prey (e.g., eagles)
G. badgers
H. rats
I. frogs
J. fish
K. flowers
L. weeds
M. soil micro-organisms
N. waterways (e.g., ponds & streams)
O. hedgerows
P. woodlands
Q. the beauty of the countryside
R. jobs and income in the countryside
S. human health
Q9. You said you are very concerned about [A to M, inserted from Q8].
How concerned would you be if only a small number of individual [A to
M, inserted from Q8] died, but the overall population was unaffected?
(asked only of those who were very concerned about effects on biota A –
M in Q8).
Q10. When thinking about buying food, which of these is the most
SID 5 (2/05)
Other
Don’t know
Cheaper food/Reduce cost of food
Grow faster/Higher yield/productivity
Improve quality of food
Improve taste of food
Improve healthiness of food
Kill unwanted pests
Control diseases of crops
Control plants that compete with crops
More profit
Other
None
Don’t know
Increases cost of food
Kill insects that are not pests
Kill plants that are not weeds
Kill wildlife/Damages wildlife
Contaminate waterways
Possible health risks for humans
Reduce quality of food
Reduce taste of food
Reduce healthiness of food
Other
None
Don’t know
Very concerned
Fairly concerned
Not very concerned
Not at all concerned
Don’t know
Very concerned
Fairly concerned
Not very concerned
Not at all concerned
Don’t know
Relative product price
Page 15 of 30
Negative Codes
Risk to/impact on human health
Risk to/impact on farmers’ health
Impact on/destruction of wildlife
Impact on countryside/habitats
Cost of using pesticides
Question
important and which is least important?
Choices
Food safety
Environmental safety
Q11
Assume you have a budget of
A
£30 per week
B
£50 per week
You may allocate this budget between the purchase of a food basket –
composed of bread, milk, fruit and vegetables - on one hand, and leisure
items on the other. Clearly, if you spend more money on food, you will
have less to spend on leisure items. Assume also you have made all your
other regular purchases. While shopping for food, you are faced with the
following choices – either:
1
‘Standard food basket’, costing:
A
£12 (£18 left for leisure)
B
£12 (£38 left for leisure)
C
£12 (£18 left for leisure)
D
£12 (£38 left for leisure)
E
£12 (£18 left for leisure)
F
£12 (£38 left for leisure)
2
‘Reduced pesticide use food basket’, costing:
A
£14 (£16 left for leisure)
B
£14 (£36 left for leisure)
C
£17 (£13 left for leisure)
D
£17 (£33 left for leisure)
E
£23 (£7 left for leisure)
F
£23 (£27 left for leisure)
Which product would you choose?
‘Standard food basket’
4.2
‘Reduced pesticide use food basket’
Results and discussion
4.2.1
Stakeholder focus groups
This section provides an overall summary of the stakeholder focus group discussions, as agreed by participants.
Areas where consensus could not be reached are identified by asterisks.
In general, what levels of effect (lethal or sublethal) that may be caused by normal agricultural pesticide use on crops are
acceptable and what are unacceptable?
For birds, there was a large degree of consensus amongst stakeholders that ideally there should be no adverse effects
on individuals, including behavioural effects. However, one, or a very few, dead or otherwise affected animals may be
acceptable*, while mortality of a large group (e.g., a flock) is not, irrespective of any population effects. Off-field effects in
addition to on-field effects are more serious than on-field effects alone. Any effect that leads to population decline outside
natural fluctuations at any scale is unacceptable, as are sublethal effects over a wide spatial or temporal scale, or product
persistence that restricts recovery. However, changes in populations need to be understood in the context of other ecological
and human factors that may influence population size. For example, the effects of pesticides should be considered in the
context of other farming practices, such as whether manual weeding in organic systems is as disruptive to ground-nesting
birds as the use of pesticides. There needs to be greater understanding of whether pesticides cause any additional effects
beyond the effects of other agronomic practices. This implies that population modelling is required, although this is a major
challenge given the limited toxicity data that are available. More post-authorisation monitoring is also required to check on
population predictions and projections. Many other issues also need to be taken into account for birds, such as food chain
uptake, bioaccumulation, the significance of sublethal effects (e.g., effects on behaviour that influence breeding success), and
the duration, frequency, scale and spatial distribution of impacts.
For mammals, the stakeholder consensus was similar to that for birds. However, there are potentially closer links
between effects on mammals and human health effects, so stakeholders agreed that effects on mammals may be less
acceptable to the public. Alternatively, public perception issues may be more important for mammals than for birds because
some mammals are clearly regarded as pests and are controlled by rodenticides. What therefore is the ‘natural’ population
abundance of rats and rabbits that should be the benchmark for population size?
For fishes and amphibians, again the stakeholder consensus was similar to that for birds. However, the discrete
nature of some water bodies (e.g., ponds), and differences in organism mobility mean that effects may be more severe in
aquatic habitats. In addition, fishes are definitely considered to be off-field in the UK, so there should be greater concern over
any predicted or measured effects on them. The visibility of fish kills may make them less acceptable to the public,
irrespective of any population effects or the time to population recovery. Amphibians may be more susceptible to population
impacts as they already appear to be in decline.
For invertebrates and plants, loss of species richness (i.e., the number of different species, not the number of
individuals) both in- and off-crop compared with what ‘should’ occur at a site is unacceptable, as are effects on population
viability. Off-crop organisms should receive greater protection than in-crop organisms. In general, there should be an
intolerance of off-field effects on either species richness or abundance of individuals within species. Any indirect effects on
the food chain are unacceptable (e.g., lack of invertebrate food for fish fry or bird chicks).
SID 5 (2/05)
Page 16 of 30
For plants alone, 100% eradication in-field is unacceptable, especially if it includes rare arable plants. There should
be tolerance in-field of weed species that are rare, do not compete with crops, or have no adverse effects on human health.
For invertebrates alone there should be active encouragement of beneficial species and no adverse effects on their
populations.
There should be no effects on other organisms as a result of impacts on micro-organisms. However, the intrinsic
value of microbes to humans is likely to be in their biogeochemical functions and not their species structure. Microbial
function should therefore ideally be maintained, although a brief reduction in function is acceptable. This reduction may be
extensive, e.g., for soil sterilants.* Maintenance of function needs to be established in the context of non-degraded microbial
assemblages.
When assessing acceptable recovery times for invertebrate, plant and microbial populations, there should be
recognition of differences in acceptable effects between in-field and off-field non-targets. In-crop recovery should be within
one year. Off-crop recovery period is probably of less importance than the spatial extent of any impacts (i.e., distance into
field margin), and should be minimal. Transient off-field effects may be acceptable.* There should be no general impairment
of population recovery potential either in-crop or off-crop. Acceptable and unacceptable recovery times will depend on
particular species life-histories and ecology. For example, effects from application of an herbicide in autumn when aquatic
macrophytes are already dying back may be acceptable, while similar effects in spring may be unacceptable. A range of
species of differing mobility and recovery potential would need to be studied as some (e.g., springtails) are known to recover
slowly.
Does your response depend upon the particular species affected?
For birds and mammals, in principle probably no, but it might depend on the conservation versus nuisance value of
particular species, and desired population abundances of pest (e.g., rat) and non-pest (e.g., dormouse) species. Despite this
distinction, vertebrate pest control should be deliberate and not an indirect artefact of insecticide, herbicide or fungicide use.
Some birds and mammals may have important ecosystem functions, while others are particularly attractive to humans, and
this may influence scientific or public concerns. Some consideration should also be given to adverse effects on the wider
genetic pool of a species. Some apparently similar species have very different reactions to exposure to some pesticides (e.g.,
Canada versus Branta geese), so toxicity data need to be extrapolated between species with care.
For fishes and amphibians, the stakeholder consensus was similar to that for birds. However, for fishes the level of
concern may also depend upon the commercial value of particular species. For example, the angling community have
significant views about the range and abundance of fish species that they expect at particular sites.
For invertebrates and plants, in general all species should be treated equally, but the answer may at times depend on
the conservation versus nuisance value of a particular species, and its overall role within the wider assemblage of organisms.
It is also likely that there will be more public support for aesthetically appealing species, such as butterflies, dragonflies and
poppies.
For microbes, species composition is probably not as important as function. However, this issue is not well studied
and considerations of microbial biodiversity may increase in importance as research and understanding increase. For
example, more study is required to determine whether application of microbial insecticides (e.g., Bti) alters natural microbial
structure in soils.
Stakeholders considered that there should be no major difference in approach when considering aquatic and
terrestrial species, except that aquatic species are always off-crop in the UK and may require more protection because of their
small-scale habitat use and restricted opportunities for recolonisation (e.g., for ponds). For aquatic systems, there may also be
a useful link between what is classified as ‘good quality’ under the Water Framework Directive, and what is acceptable to the
public.
What environmental scale does your response relate to (e.g., a single field, a parish, a county, or the whole country)?
For birds, whatever scale is necessary to sustain a particular species population, given its range, territory and
mobility. In general this translates into a field scale for kills, and county to regional scale for population effects. However,
there may be more local issues such as the desire for songbirds in gardens, which will be influenced by the territoriality and
life history strategies of certain species. This question also needs to be addressed in the context of other population stressors
and the spatial structure of the agricultural landscape. The same treatment applied to a mixed farming landscape is likely to
have a lower impact than when applied to a crop monoculture.
For mammals, fishes and amphibians, mobility is likely to be more restricted than for birds, so protection should be
at a more local scale, such as the field or individual water body, to sustain particular species populations. Most mammals,
fishes and amphibians also do not move into cropped areas in the same way as birds, so spatial considerations may differ
substantially.
For invertebrates, plants and microbes, the field scale is important for loss of richness (this may be more important
for organisms such as plants, with localised populations). Some stakeholder groups suggested county to regional scales as
important for population effects. Other groups suggested that the scale should be related to particular species needs and the
patchiness of impacts in their habitats, and there was a suggestion that an operational scale might be 25-50km2 grids. It was
also pointed out that, in general and if possible, this issue should be looked at within a wider landscape use and metapopulation context. For example, putting an entire parish under oilseed rape will have ecological consequences irrespective
of pesticide use. More diverse agriculture with consequent ‘patchy’ application of pesticides is likely to be less detrimental.
SID 5 (2/05)
Page 17 of 30
What frequency of occurrence does your response relate to (e.g., acceptable once a year, once every 10 years, etc.)?
For vertebrates, stakeholders felt that an occasional local occurrence of effects on a small number of individual birds
and mammals, but not fishes or amphibians, is probably acceptable, so long as they are not protected species.* Interactions
with the timing of migration (birds) and hibernation (mammals) also need consideration. No frequency of occurrence leading
to population decline over the long term (e.g., 20-30 years, and taking natural stressors into account) is acceptable. Responses
to stressors depend upon the resilience and recovery rates of different populations, and will relate to interactions between
generation time and fecundity and the direct and indirect effects of pesticides. It should be possible to make projections of
these with population models to aid decision-making. However, appropriate benchmarks need first to be defined for what
population sizes for different species ‘should’ occur in farmland (which is itself a human artefact), and efforts made to
achieve this. Small effects every year in the same place or across widespread parts of the UK are less acceptable than larger
effects at very infrequent intervals in particular places. A mechanism is required for rationally balancing the benefits and
essential uses of a pesticide against any environmental effects.
For invertebrates, plants and microbes, large repeated in-crop effects each year are acceptable so long as recovery
occurs by the beginning of the following season. For the off-crop environment an occasional local occurrence of effects on a
small number of individuals is acceptable, so long as they are not protected species. No frequency of occurrence leading to
widespread population decline is acceptable. Again, it was suggested that this should be looked at within a landscape context.
The key issue is what frequency of effects across the landscape is compatible with the maintenance of populations and food
chains across the same landscape.
Conclusions
Stakeholders in focus groups, who constitute a cross-section of UK expertise on pesticide effects, were almost
entirely concerned with the potential effects of pesticides on animal and plant population viability, and micro-organism
function. Focus group discussions returned regularly to this theme, so that questions about acceptable geographical and
temporal frequencies of effects and recovery times were answered in terms of organism life history strategies and the ways in
which particular frequencies of impacts might affect population demographics. This is not to suggest that stakeholders were
entirely unconcerned about individual organism welfare. However, all stakeholder groups recognised that a trade-off exists
between the potential economic advantages of responsible pesticide use and the potential disadvantages of individual
poisoning events. So long as individual off-field poisonings are infrequent and do not adversely affect population size and
viability, stakeholders generally felt that the use of a pesticide leading to these effects should be considered as acceptable.
This view is in keeping with guidance provided in support of Directive 91/414/EC (Campbell et al. 1999, EC 2002a&b,
Giddings et al. 2002), which focuses on protecting populations rather than individual organisms.
4.2.2
Public opinion survey
Products used at home or in the garden
Most respondents (71%) had used at least one pesticide product in the home or garden.
Important factors that society should take into account in deciding how to grow food
Respondents were more concerned about the effects of food-growing methods on human health (65%) and food
quality (52%) than on wildlife health and variety (32%), animal welfare (24%) or waterways (13%).
Possible causes of damage to wildlife and habitats in Britain
Pesticides were the single possible cause of damage to wildlife and habitats most frequently identified by
respondents (26%), with the other possible causes mentioned on the questionnaire identified as such by 15% or less of
respondents. Nineteen percent of respondents had not heard of any of the identified factors as possible causes of damage.
Support for use of pesticides on farm crops
More respondents opposed (31%) or strongly opposed (15%) the use of pesticides on farm crops than supported
(19%) or strongly supported (1%) this use. Around a third of respondents had no strong views either way. When respondents
who supported the use of pesticides (n=419) were asked why they held this view, a third or more agreed that it was because
pests were killed or diseases were controlled by the use of pesticides. Twenty-five percent of these respondents also
expressed the view that the amount of food grown could be increased through pesticide use. When respondents who opposed
the use of pesticides (n=969) were asked why they held this view, most (65%) agreed that it was because of risks to human
health, with 42% expressing the view that it was because of risks to wildlife. Similar views emerged when respondents were
asked about the benefits and costs of pesticide use.
Concern for particular environmental features
Figures 1a and 1b show the concerns of respondents for the effects on particular environmental features of either i)
non-specific human activities or ii) pesticides. In general, respondents ranked their concerns in a similar order, with effects
on human health regarded as most important, and effects on highly regarded wildlife such as songbirds and badgers also
ranking relatively highly. In contrast, effects on relatively unloved organisms such as rats and weeds were not ranked as
important.
However, there were some interesting differences in the views of respondents asked about generic human activities
and those asked specifically about pesticides, which are summarised in Figure 1c. More respondents were very concerned
about the effects of pesticides, when compared with the effects of general human activities, on waterways, human health,
SID 5 (2/05)
Page 18 of 30
fishes, soil micro-organisms, frogs, earthworms, carrion birds, butterflies, insects and badgers. In contrast, respondents
tended to be less concerned about the effects of pesticides when compared with general human activities on woodlands,
countryside jobs and income, and countryside beauty.
Respondents who were ‘very concerned’ about the effects of human activities or pesticides on particular organisms
were then asked what their view would be if only a small number of individuals died, but the overall population remained the
same. In both cases, between 40% and 55% of respondents did not change their views and remained very concerned.
However, the remaining respondents did revise their opinion and expressed less concern. A greater percentage of respondents
remained very concerned about the effects of pesticides when compared with the effects of general human activities on
insects, weeds, butterflies, songbirds, birds of prey, earthworms, carrion birds and badgers.
Economic considerations
When buying food, respondents ranked consideration of food safety as more important than product price, which
was in turn considered more important than environmental safety. This ordering changed when respondents considered the
least important considerations when buying food, with product price less important than environmental safety, which was
less important than food safety. Although this suggests that price was consistently of low importance to respondents, there
were clear trade-offs when the relative price of a ‘no-pesticide’ food basket and the available budget were varied (Figure 2).
When the additional cost of buying non-pesticide food was small, half of the respondents chose a food basket produced
without pesticides, and there was little difference in this choice when the total available budget was £30 or £50. As the
additional cost of a ‘no pesticide’ food basket increased, fewer respondents were prepared to choose it, particularly when the
overall budget was relatively low.
A conditional logit model was fitted to these data and parameters for the linear indirect utility function were
estimated, enabling calculation of willingness to pay (Table 3), conditioned on gender, age (below or above 45 years old),
weekly food and leisure budget, and primary concern while buying food. Results broadly conformed to prior expectations.
Respondents faced with a higher budget were willing to pay more, but the fact that this increase was not very large suggests
that consumers do take into account substitutes (in this case leisure), and therefore the use of substitute reminders is
warranted.
Table 3 Willingness to pay estimates
Gender
Male
Female
Age
Younger
Older
Younger
Older
£30 per week budget
Relative
Food
price
safety
2.6
14.9
5.4
17.2
7.4
26.2
10.4
27.5
Environmental safety
14.3
15.6
19.3
20.4
£50 per week budget
Relative
Food
price
safety
4.4
16.9
7.0
18.9
9.5
28.3
12.1
29.3
Environmental safety
15.4
16.6
20.4
21.4
Other factors
The information presented in this report is a complete summary of the results from the stakeholder focus groups, but
summarises only the headline results from the survey of public opinions. The survey opinions can also be stratified by
gender, age, social class, household income, presence of children in the household, geographic region, rurality, and level of
education. For example, female respondents were willing to pay more than male respondents for a no-pesticide food basket,
as were older when compared to younger respondents. Detailed stratified results are not presented here, for reasons of space,
but are freely available by email from Mark Crane (craneconsultants@aol.com).
4.2.2
Conclusions and Relevance for Pesticide Risk Management
The survey of opinions from more than 2000 members of the public in the present study showed that although
pesticides are widely used in homes and gardens, use on farm crops remains of concern to the public. Concerns are greatest
on issues of human health and food quality, as noted in other surveys of public attitudes on pesticides (Anon 2000, Dunlap
and Beus 1992). However, potential environmental effects are also an issue for a substantial number of people, particularly if
attractive species could be affected. In contrast to the stakeholders in focus groups, about half of the public would remain
very concerned about the effects of pesticides even if they affected only individual organisms and not populations.
Despite these apparently widespread concerns over pesticide use on farm corps, the desire to purchase food that has
not been produced with the use of pesticides depends on its price relative to the total available budget, with only one-fifth of
the public, or fewer, prepared to pay a substantial additional amount for this. This identifies a divergence between the
public’s perceptions of the potential risks associated with pesticides, and their practical willingness to pay for food
production that involves none of these potential risks. Such findings are in agreement with Tait et al. (2001) who found from
a review of the literature on chemicals and public values that there is usually only a rather weak correlation between public
attitudes and values, on one hand, and actual behaviour on the other, often because of intervening variables, such as price.
Female respondents were willing to pay more than male respondents, which agrees with Loureiro et al. (2002) who, in a
similar context, estimated willingness to pay for eco-labelled apples, and to Veeman and Adamowicz (2000) in the case of
willingness to pay for ‘safer’ skimmed milk. The latter suggest that women are more likely to purchase the majority of
household food items and are therefore more aware of food safety issues. Also in agreement with Veeman and Adamowicz
(2000) older respondents were willing to pay more than younger respondents. Finally, respondents with food safety and
environmental safety as their main food shopping concern had a higher willingness to pay, while price-sensitive respondents
expressed a lower willingness to pay. Again, these results agree with Loureiro et al (2002).
SID 5 (2/05)
Page 19 of 30
Previous studies have shown that experts and the public tend to rank the relative generic risks from pesticides rather
consistently. For example, Slovic (2000) asked experts and lay people to rank the perceived risks of 30 potentially hazardous
activities. In his survey, the lay people ranked ‘Pesticides’ ninth, while the experts ranked them eighth. This convergence is
in marked contrast to activities associated with significant public dread (Perrow 1999), such as ‘nuclear power’ which was
ranked as the most important hazard by lay people but was ranked only 20th by experts. In contrast to these findings, the
current study suggests that when compared with experts in stakeholder groups the public believes that there are greater
environmental risks from pesticide use, possibly because more ‘dreadful’ threats such as nuclear accidents were not included
in the survey.
Many scientists and industrialists believe that greater public understanding of science is the solution to public
attitudes that seem to be irrational, or are at variance with expert views or the actual behaviour of the public. However, social
science studies show that this is not the solution because once a person’s mind is made up about fundamental values they will
use only the scientific information that supports their position, ignoring the science that does not (Tait 2001b)
a
Very concerned
Fairly concerned
Not very concerned
Not at all concerned
Don't know
60%
Percentage of respondents
50%
40%
30%
20%
10%
Ra
ts
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Environmental feature
b
Very concerned
Fairly concerned
Not very concerned
Not at all concerned
Don't know
60%
Percentage of respondents
50%
40%
30%
20%
10%
Environmental feature
SID 5 (2/05)
Page 20 of 30
R
at
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15%
Very concerned
Fairly concerned
Not very concerned
Not at all concerned
Don't know
5%
0%
Countryside beauty
Countryside jobs & income
Woodlands
Hedgerows
Rats
Birds of prey
Weeds
Flowers
Songbirds
Badgers
Insects
Butterflies
Carrion birds
Earthworms
Frogs
Soil micro-organisms
-10%
Fishes
Human health
-5%
Waterways
Percentage difference in response
10%
-15%
Environmental feature
Figure 1
Concerns of respondents over the effects of a) human activities in the countryside and b) pesticides
on different environmental features; c) differences in the concerns of respondents over the effects of
human activities versus the effects of pesticides in the countryside on different environmental
features. Positive values mean that respondents were more concerned about pesticides.
Difference in percentage of respondents choosing 'non-pesticide'
food basket over 'pesticide' food basket
60%
51%
49%
50%
38%
40%
30%
27%
20%
20%
16%
10%
0%
£2 (total budget
available = £30)
£2 (total budget
available = £50)
£5 (total budget
available = £30)
£5 (total budget
available = £50)
£11 (total budget
available = £30)
£11 (total budget
available = £50)
Additional cost of 'non-pesticide' food basket and total budget available (£)
Figure 2
5.
Effect of price and available budget on respondents’ choice of food basket produced with or without
pesticides
DECISION FRAMEWORK
5.1 Issues to consider when constructing a decision framework for pesticide authorisation
Earlier outputs from this project show that the following ecological and socioeconomic issues should, where possible,
be considered when constructing a decision framework for authorising pesticides:
1. Demographic data from laboratory studies should be applied more effectively during pre-authorisation pesticide
risk assessment.
2. Production of a formal conceptual ecological risk assessment model for each product or active substance for which
authorization is sought would provide an appropriate framework for integrating and applying ecological
knowledge.
3. The current lack of targeted post-authorisation monitoring for the effects of pesticides in the natural environment
leaves great uncertainty over the level of precaution of pre-authorisation risk assessment and justifies conservatism
during this assessment.
4. Long-term adverse pesticide effects on populations of plants and animals, or on microbial function, are considered
unacceptable by all stakeholders and the public, and should be assessed within the geographical and temporal
context of individual species life histories.
SID 5 (2/05)
Page 21 of 30
5.
Individual vertebrate deaths should be avoided, if at all possible, even if these deaths are unlikely to affect
population size or viability, and particularly if the organisms at risk are those that the public care about.
In addition to this, a useful decision framework must also be compliant with current regulatory requirements and be
based upon scientific evidence and knowledge, except when decisions need to be made in the absence of such evidence or
knowledge.
5.2 Proposed decision framework
The proposed decision framework operates on the principle that scientific evidence, when available, takes precedence
over the opinions of stakeholders and the public and that the trigger values stipulated by Directive 91/414 must be observed
during the authorisation process. The framework therefore changes nothing during the earlier tiers of pesticide risk
assessment. However, it does offer improvements at higher tiers when judgement must be applied in determining whether a
predicted level of environmental effect is acceptable or not. It does this by weighting the importance of each result according
to the value placed upon it by stakeholders and the public. The main contribution to this from the stakeholder groups is the
consensus that the population viability of all non-target higher organisms, and the community function of microbes, must be
maintained under all circumstances. This is entirely consistent with current practice in pesticide risk assessment and
guidance.
The results from the public opinion survey are then used to determine weightings if there are predicted effects on
individual organisms, or if the results from higher tier mesocosm or field studies suggest that transient effects may occur. The
weightings are based on survey respondents who were either ‘very’ or ‘fairly’ concerned about the effects of pesticides on a
particular group of organisms, multiplied by the proportion of very concerned respondent who remained very concerned even
when told that only a few individuals would be affected (Questions 8 and 9 in Table 2 above). The weightings derived from
this are shown in Table 4. We should like to emphasise that if experts truly have scientific knowledge that can be applied
consistently to interpret effects observed in such studies then this should be used in preference to use of public opinions.
However, if there is no quantifiable expertise on the issue under discussion then we would argue that the public’s values
should take precedence. This could be done as follows:
1. List the receptors in the environment of concern (in this case aquatic ecosystems). If the pathways or interactions
between receptors are complex, then a conceptual model is likely to be the best way to capture and visualise all
appropriate information, as discussed in Section 3 of this report.
2. Calculate a maximum achievable score that could be achieved for this system if all hazard quotients or TERs are 1,
weighted by the overall weights in the final column of Table 4.
3. Calculate the actual score for the system by substituting the true TER, if it is less than 1. No concessions are made
for receptors for which TERs are very high (i.e., very low risks) because this would counterbalance and therefore
mask receptors for which TERs are low, so a fixed TER of 1 is applied if the measured TER is >1.
4. Calculate the ratio between the maximum and actual scores and compare with a pre-determined pass or fail
threshold. For example, this could be that the sum of all the individual receptor scores must be no less than 90% of
the total maximum score, and that no individual score for a receptor can be less than 50% of the maximum score
for that receptor. These threshold values may be set more formally by using empirical information from past
decisions made during the authorisation process.
Expert judgement can be used to challenge the results from this formal decision making process on the basis of
additional scientific evidence, but this would be the exception rather than the rule. The approach is illustrated in a short,
simplified case study in the next section.
Table 4 Weights for incorporating public values into the proposed decision framework. Weights for organism groups
‘in general’ are mean values.
Organism
Proportion of respondents very or
fairly concerned about pesticide
effects (A)
Songbirds
Carrion birds
Birds of prey
0.82
0.63
0.79
Proportion of respondents
remaining very concerned when
effects on only a few individuals
(B)
0.51
0.46
0.5
0.42
0.29
0.40
0.74
0.21
0.46
0.49
0.34
0.10
0.8
0.66
0.48
0.53
0.38
0.35
0.74
0.59
0.54
0.5
0.4
0.30
0.57
0.73
0.3
0.5
0.44
0.52
0.29
0.32
0.16
Overall weight
(A*B)
Birds in general
Badgers
Rats
0.37
Mammals in general
Fishes
Frogs
0.22
Aquatic vertebrates in general
Butterflies
Insects
0.37
Arthropods in general
Earthworms
Flowers
Weeds
0.35
Plants in general
SID 5 (2/05)
0.24
Page 22 of 30
5.3 Decision framework case study
Table 5 shows some hypothetical data for Pesticide X, an organophosphate insecticide. Briefly, an assessment of single
species data for this substance would find the following, on the basis of TER trigger values:
1. Negligible risks to mammals, fishes, algae, earthworms, plants and micro-organisms.
2. Marginal acute, but not chronic, risks to birds.
3. Considerable acute risks to aquatic and terrestrial arthropods.
These single species data are likely to prompt an aquatic mesocosm study, the results of which are also presented, and
which suggest that potential risks remain for aquatic arthropods at estimated exposure levels. The EC guidance document
(EC 2002a) suggests that expert judgement should be used at this stage, although it is difficult to see what knowledge, rather
than opinion, an expert could bring to bear on this issue. This is because personal values rather than any scientific knowledge
are likely to determine whether an expert decides that the critical value from the mesocosm study is the community NOEC,
the concentration that causes only temporary effects on arthropods, or the concentration at which more persistent effects are
evident on a few arthropod species.
Table 5. Data for a hypothetical organophosphate insecticide (Pesticide X)
Test organism/system
Mallard
duck
(Anas
platyrhynchos)
Mallard
duck
(Anas
platyrhynchos)
Rat (Rattus norvegicus)
Rainbow trout (Oncorhynchus
mykiss)
Waterflea (Daphnia magna)
Green
alga
(Pseudokirchneriella
subcapitata)
Duckweed (Lemna minor)
Honeybee (Apis mellifera)
Earthworm
(Lumbricus
terrestris)
Soil micro-organisms
Sewage micro-organisms
Aquatic mesocosm
Toxicity
LD50 980 mg/kg
Maximum Estimated
Theoretical Exposure
120 mg/kg
Toxicity:Exposure Ratio
(TER)
8.2
Annex VI TER trigger
and pass/fail result
10 (fail)
Reproduction NOEC
640 mg/kg
LD50 >1300 mg/kg
LC50 115 µg/L
120 mg/kg
5.3
5 (pass)
120 mg/kg
0.9 µg/L
>10.8
127.8
10 (pass)
100 (pass)
EC50 0.01 µg/L
EC50 1500 µg/L
0.9 µg/L
0.9 µg/L
0.01
1667
100 (fail)
10 (pass)
EC50 700 µg/L
LD50 0.2 µg/bee
LC50 160 mg/kg soil
0.9
960 g/ha
2.5 mg/kg soil
778
4800
64
10 (pass)
Hazard Quotient of 50 (fail)
10 (pass)
NOEC for nitrogen cycling
32 mg/kg soil
LD50 280 mg/L
Community NOEC 0.5
µg/L.
Effects
on
arthropod
abundance for six weeks at
1.0 µg/L.
More persistent effects (>8
weeks) on abundance of
some arthropod species at
3.0 µg/L
2.5 mg/kg soil
12.8
10 (pass)
0.9 µg/L
0.7 µg/L
311111
Community NOEC:
0.7
Effects for six weeks: 1.4
Persistent effects: 4.3
10 (pass)
No set trigger value. If
PEC:PNEC ratio of 1 is
acceptable threshold then
Community NOEC falls
below this value.
If, instead, we follow the simple proposed decision framework for Pesticide X (Table 6), we begin by listing the main
receptor groups in the system within which potential risks may occur (aquatic systems). We then assign maximum achievable
TER scores of 1 to each receptor in this system, weighted by the extent to which the public is concerned about effects on
individuals from each receptor group, or organisms that are similar to these receptors. The actual TER for each receptor is
then input, and the two sets of scores are compared. In this very simple example, pesticide X would be authorised because the
overall acceptability score is above the threshold of 90% and no single receptor has an acceptability score below the
threshold of 50% (assuming that these thresholds are agreed as reasonable). It is clear from this example that the decision
framework proposed here will only be useful when decisions are marginal, because a TER of <0.5 (i.e. predicted exposure is
twice the no effect value) for any receptor will automatically produce an acceptability score of <50%, leading to rejection.
However, it is precisely this marginal area that is important, as correct decisions will usually be self-evident to risk managers
when TERs are either low or high for individual receptors or overall.
Table 6 Example for pesticide X of a formal decision framework incorporating public values and TERs
Receptor
Waterfowl
Aquatic mammals
Fishes
Arthropods
Other invertebrates
arthropods)
Algae
Aquatic macrophytes
SID 5 (2/05)
TERactual (=1 if
TER > 1)
Maximum achievable
acceptability score
(Public value x TER1)
(non-
0.37
0.22
0.38
0.35
0.29
1
1
1
0.7
1
Actual
acceptability
score (Public
value x TERactual
0.37
0.22
0.38
0.25
0.29
0.16
0.32
1
1
0.16
0.32
Page 23 of 30
Ratio
actual:maximum
acceptability score
1
1
1
0.71
1
1
1
Receptor
Maximum achievable
acceptability score
(Public value x TER1)
TOTAL
TERactual (=1 if
TER > 1)
2.09
Actual
acceptability
score (Public
value x TERactual
1.99
Ratio
actual:maximum
acceptability score
0.952
The method described here is generally consistent in that,
1. Higher scores represent a utility that corresponds to the wishes expressed by the public.
2. The scores allow for discrimination between combinations of potential events.
3. There is a maximum achievable score that corresponds to a ‘perfect’ score.
4. All pesticides can be measured and assessed on the same scale.
6.
7.
8.
MAIN FINDINGS AND IMPLICATIONS
1.
There is currently little information available to determine whether normal use of pesticides poses long-term risks
to wildlife. There is clearly the potential for risk, but the temporal and spatial realisations of this risk in the natural
environment remain largely unknown. Reliable population projections of pesticide effects during pre-authorisation
risk assessment require reporting of relevant demographic data (age-specific survival and reproduction), which is
not currently the case. Post-authorisation monitoring for effects is also sparse, but is the only effective means of
testing the effectiveness of the authorisation process. Pesticide risk assessments would also benefit from formal
adoption of conceptual modelling approaches, which are widely used in other ecological risk assessment
frameworks. These shortcomings in knowledge and procedure mean that current pesticide risk assessments should
remain conservative in their assumptions.
2.
Stakeholders in pesticide risk management from across the spectrum of views agree that maintenance of higher
organism population viability and microbial function are the main aims of pesticide risk assessment and
management. Substantial numbers of the general public are more conservative than these informed stakeholders,
but a softening of their views is evident when only a few individual organisms, rather than entire populations, are
potentially affected by pesticides,. Despite the public’s apparent concerns over pesticides, economic factors clearly
influence the extent to which they are prepared to pay for food that has been grown without the use of pesticides. A
balanced approach to incorporating public values into pesticide risk management would therefore take account of
the public’s views on potential costs (i.e., environmental effects, or effects on human health) and the potential
benefits of pesticides, although this is not currently possible under existing regulations.
3.
A simple decision framework for pesticide risk management can be constructed which takes public values into
account when interpretation of environmental information is value-laden rather than evidence-based. This
framework complies with Directive 91/414 and its associated guidance and provides a consistent means for
deciding on the acceptability of observed effects.
REQUIREMENTS FOR FURTHER RESEARCH
1.
There is a clear need for targeted monitoring studies in aquatic and terrestrial environments to determine the spatial
and temporal patterns of any adverse effects caused by the approved use of pesticides.
2.
The public should be surveyed at regular intervals to determine the values that they place upon different
environmental receptors. This survey is most cost-effectively achieved as part of an omnibus survey performed by a
professional polling organisation. The survey should attempt to obtain information on a broader range of organisms
than in the current project so that there is less need to extrapolate public values between taxonomic groups.
3.
The proposed decision framework should be applied to a wide variety of currently authorised substances to
determine whether it is fit for purpose and provides improved consistency of decision-making. Three main areas
should be explored:
a. whether it is possible to construct generic conceptual models for aquatic and terrestrial environments which
capture all of the main receptors that may be directly or indirectly affected by pesticide exposure, as
suggested by Objective 1 of this project;
b. whether the information supplied by notifiers in support of authorisation is sufficient to parameterise these
generic conceptual models; and
c. whether quantitative thresholds for acceptable ‘scores’ from the decision framework can be determined
from the empirical data.
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References to published material
9.
This section should be used to record links (hypertext links where possible) or references to other
published material generated by, or relating to this project.
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9.1 Refereed journal papers
Crane M, Norton A, Leaman J, Chalak A, Bailey A, Yoxon M, Smith J, Fenlon J. Acceptability of pesticide impacts
on the environment: what do United Kingdom stakeholders and the public value? Submitted to Pest Man
Sci, April 2005.
Newman MC, Holloway GJ, Crane M. Current pesticide risk assessment in Europe: does it account for long-term
adverse effects on non-target organisms? Submitted to Rev Environ Contam Toxicol, June 2005.
Crane M, Fenlon J, Smith J. A consistent environmental decision framework for pesticide risk management. In
preparation.
9.2 Conference presentations
Crane M, Norton A, Leaman J, Chalak A, Bailey A, Yoxon M, Smith J, Fenlon J. Public acceptability of pesticide
effects. SETAC Europe 15th Annual Meeting, Lille, France, 22-26 May 2005.
Newman MC, Holloway GJ, Crane M. Current pesticide risk assessment in Europe: does it account for long-term
adverse effects on non-target organisms? SETAC Europe 15th Annual Meeting, Lille, France, 22-26 May
2005.
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