POLLINATORS POLLINATION AND FOOD PRODUCTION

THE ASSESSMENT REPORT ON POLLINATORS, POLLINATION AND FOOD PRODUCTION
in nectar and pollen. In addition, honey bees may also
be exposed to beekeeper-applied treatments such as
antibiotics and varroacides (Chauzat et al., 2009; Mullin
et al., 2010) There is evidence of multiple residues of
pesticides detected in bees, honey, pollen and wax within
honey bee colonies (e.g. Thompson, 2012) but these data
are complex in terms of the number, scale and variability
of pesticide residues. Data are very limited or absent for
other pollinators and for the effects of complex pesticide
mixtures.
There is strong evidence that when combinations of
pesticides have been screened in a range of aquatic
invertebrates (Verbruggen and van den Brink, 2010;
Cedergreen, 2014), synergistic interactions (resulting in
greater than 2-fold increase in toxicity when compared
BOX 2.3.5
Assessing the possible contribution of neonicotinoids to pollinator declines: What do we still need to know?
To date the role of neonicotinoids in pollinator declines has
been a particularly polarised debate. There are both qualitative
and quantitative aspects, so what evidence do we need to
inform the debate?
Where declines in species and possible drivers have been
identified but not prioritised, we need to weigh the evidence
carefully, and identify which are the key gaps (e.g. (Van der Sluijs
et al., 2013; Godfray et al., 2014; Lundin et al., 2015). Where
the evidence is still scant, Hill’s epidemiological criteria can be
used to identify whether the logic criteria (coherence, plausibility,
gradient) coincide with the circumstantial epidemiological
evidence, e.g. for honey bee declines (Cresswell et al.,
2012a; Staveley et al., 2014). Such an analysis both identifies
knowledge gaps, but also helps to differentiate between
the differing drivers of declines. For example declines of
bumble bees in the 1950s were certainly not initiated by
neonicotinoids, but probably due to loss of flower-rich habitat
with agricultural intensification (Ollerton et al., 2014).
Apart from dust generated during drilling of treated seed or
off-label applications, national incident monitoring schemes
suggest approved neonicotinoid use has not been associated
with honey bee mortality. However, vigilance is needed to
ensure that approved uses include mitigations to protect
pollinators and the environment (e.g. buffer zones to off-crop
areas, not applying to bee-attractive crops in flower or crops
containing flowering weeds) and that use instructions are clear,
understood and respected. Concerns have arisen primarily
from acute or chronic sub-lethal exposures that might interfere
with foraging, orientation and learning abilities and other
behavioural characteristics of pollinators, as well as with the
immune system at the individual and colony level.
There remain some key gaps in our knowledge:
1. Toxicity. There are large differences in the toxicity of
neonicotinoids in honey bees, e.g. thiacloprid and acetamiprid
vs. imidacloprid, clothianidin and thiamethoxam as well as
their metabolites (Blacquière et al., 2012). Although, with
appropriate assessment factors, acute (lethal) toxicity data
for honey bees can be used as a surrogate for other species
(Hardstone and Scott, 2010; Arena and Sgolastra, 2014),
large differences in species sensitivity may occur (as for
other invertebrates, e.g. Cloen (Mayfly) compared to Daphnia
(Roessink et al., 2013)). The ability of bees to detoxify and
excrete ingested neonicotinoid residues contributes to species
differences in their chronic sensitivity (Cresswell et al., 2012b;
Laycock et al., 2012; Cresswell et al., 2014). Therefore further
data are required especially for wild pollinator species, to
confirm that extrapolation between species is appropriate for
neonicotinoids and their metabolites (Lundin et al., 2015).
Even less is known about sub-lethal toxicity, e.g. at which
doses are no effects found, which effects are important for
which species (see Figures 2.3.5-2.3.7 (Lundin et al., 2015))?
For example, there is a plausible potential for interactions
between sub-lethal exposure to neonicotinoids and foraging
efficiency, resulting in effects at the colony level for species
with low numbers of foragers (Rundlöf et al., 2015). The
Rundlöf et al. (2015) study showed that, whilst there were no
effects on honey bee colonies, exposure to flowering springsown
oilseed rape grown from seed treated with the highest
approved application rate of clothianidin in Sweden affected
bumble bee colony development, Osmia nest establishment
and the abundance of wild bees observed foraging on the
crop. The residue levels in pollen and nectar were higher than
previously reported in oilseed rape (Blacquière et al., 2012;
Cutler et al., 2014a; Godfray et al., 2014) and highlight the
need for understanding of the variability of pesticide residue
levels in crops. For example, in Europe, varieties of oilseed
rape sown in the autumn/winter are far more prevalent than
spring-sown varieties. Autumn/winter sown varieties are
often treated with lower levels of neonicotinoid and the time
from sowing to flowering is about 7-8 months, rather than
3-4 months for spring varieties. However, these results are of
considerable importance, because they show for the first time
the effects under field conditions of a neonicotinoid insecticide
on wild bees in the absence of an effect on honey bees. In
order to quantify the possible contribution of these sublethal
effects to the observed declines we need not only to test at
levels that result in these effects under laboratory conditions,
(Figure 2.3.5) but also at field-realistic exposure levels and
profiles (Lundin et al., 2015). Such an approach may use
designs similar to that of Rundlöf et al. (2015) to evaluate the
effects on managed and wild bee populations of the most
widely used insecticides, applied according to their approved
use, in the most widely grown pollinator attractive crops.
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2. DRIVERS OF CHANGE OF POLLINATORS,
POLLINATION NETWORKS AND POLLINATION