POLLINATORS POLLINATION AND FOOD PRODUCTION

THE ASSESSMENT REPORT ON POLLINATORS, POLLINATION AND FOOD PRODUCTION
58
2. DRIVERS OF CHANGE OF POLLINATORS,
POLLINATION NETWORKS AND POLLINATION
exposure for honey bees, although data on water sources
are more limited for other bee species (Pistorius et al., 2012;
Godfray et al., 2014). Another potential route of exposure
is the generation of dust, containing insecticide, that may
drift onto nearby flowering crops or weeds during drilling
of treated seed (Krupke et al., 2012; Pisa et al., 2014).
There have been a number of studies demonstrating the
lethal effects of dusts generated from neonicotinoid-treated
seeds during drilling (Bonmatin et al., 2015) and large-scale
honey bee mortality has resulted from treated seed when
the seed contained high levels of dust particularly when
it was incorrectly coated or dust based seed lubricants
were added during drilling when dust drifted onto flowering
crops and weeds (Pistorius et al., 2009; PMRA, 2014).
There is evidence that appropriate technical measures can
be adopted to reduce the associated risk of dust although
no single measure has currently been shown to be totally
effective (Kubiak et al., 2012; Nuyttens et al., 2013).
There is evidence that the identity of pesticides present
and scale of the exposure of honey bee colonies (levels
in pollen, nectar/honey and wax) differ between crop type
(Pettis et al., 2013) and regions reflecting differences in
pesticide approval and use (Bogdanov, 2006; Johnson et
al., 2010; Mullin et al., 2010; Chauzat et al., 2011; Al-Waili
et al., 2012). However, quantitative data on an individual
pollinator’s exposure to pesticides is limited, i.e. actual
ingestion by a foraging bee, not measured residues. Pollen
and nectar consumption has been almost entirely studied in
honey bees and often extrapolated from estimated nutritional
requirements as a proxy for foraging rate (Thompson, 2012)
rather than measured directly. Exposure factors have been
evaluated for wild bees on focal crops in Brazil, Kenya and
the Netherlands by (van der Valk et al., 2013). The overall
likelihood of exposure of wild bees to pesticides were
evaluated as “probably similar” to Apis mellifera in the case
of Apis mellifera scutellata and Xylocopa, but due to a lack
of information were “unclear” for Patellapis and Megachile
and “possibly greater” for Halictidae. However, from a review
of the literature it is clear there is a lack of accurate data on
key aspects of the biology of non-Apis species (e.g. nectar
consumption by foraging bees) to allow exposure under field
conditions to be quantified.
Pesticides may result in impacts on pollinators without direct
exposure. Indirect effects on pollinators include the removal
of nectar/pollen sources and/or nest sites by herbicides
(Potts et al., 2010). Together both direct and indirect
effects of pesticides, in combination with other aspects of
monoculture agriculture, may contribute to observations
at the landscape scale of a tendency for reduced wild bee
and butterfly species richness in response to pesticide
application (Brittain et al., 2010; Brittain and Potts, 2011;
Vanbergen et al., 2013).
2.3.1.3 Evidence of lethal effects during
pesticide use
Insecticides vary widely (several orders of magnitude) in
toxicity to pollinators depending on their mode of action
(see Table 2.3.2) and target life-stage (e.g. insect growth
regulators only directly affect larvae/pupae). Even within an
insecticide class, toxicity can vary from a few nanograms
(ng) per bee to several thousand micrograms (µg) per bee,
as in the case of the neonicotinoids (Blacquière et al.,
2012). There is evidence that the detoxification enzymes in
honey bees are less diverse than in other insects making
them less well adapted to respond to exposure to a range
of chemicals (Johnson et al., 2010; Mao et al., 2013) and
even this limited range of enzymes is also affected by the
age of the bee, the time of year, etc. (Smirle and Winston,
1987). However, there is also evidence that Apis mellifera is
no more sensitive to insecticides than other insect species
(Hardstone and Scott, 2010). The relative sensitivity of
different bee species to the acute (single exposure) effects
of insecticides and other pesticides is similar, i.e., the acute
toxicity (LD 50
) is within an order of magnitude (Arena and
Sgolastra, 2014), particularly if body mass (80-300mg) is
taken into account (Arena and Sgolastra, 2014; Fischer
and Moriarty, 2014). However, the chronic toxicity (LC 50
) of
pesticides may be more variable; some evidence suggests
clearance of insecticides may differ among species of
bees (Cresswell et al., 2014). Other factors have also
been identified as affecting the toxicity of insecticides
to honey bees, including nutrition (Godfray et al., 2014;
Schmehl et al., 2014) and disease (Vidau et al., 2011) (see
section 2.4.1).
The largest published databases on acute pesticide effects
under real-use field conditions are formal incident monitoring
schemes that are limited to honey bees (only a handful
of reported incidents have involved bumble bees). These
schemes have been instigated by national governments in a
number of European countries, Australia, Canada, USA and
Japan (OECD, 2010) and are reliant on notification of honey
bee deaths either on a voluntary basis by beekeepers or
as a requirement for pesticide registrants. A single incident
may range from a few bees to several thousand bees but
has rarely been linked to an assessment of the longer-term
impact on the colony, e.g., the neonicotinoid seed treatment
dust incident in Germany (Wurfel, 2008). Where voluntary
reporting exists there is potential for under-reporting due
to reticence of beekeepers to report incidents and risk the
loss of apiary sites with good forage often on land belonging
to farmers (Fischer and Moriarty, 2014). The longestrunning
incident schemes are primarily in Europe (Germany,
Netherlands and UK), where the number of incidents where
pesticides have been identified as a cause declined from
circa 200 incidents per year in the 1980s to around 50
by 2006 (Barnett et al., 2007; Thompson and Thorbahn,
2009); more recent data from the UK show a decline from