POLLINATORS POLLINATION AND FOOD PRODUCTION

THE ASSESSMENT REPORT ON POLLINATORS, POLLINATION AND FOOD PRODUCTION
neonicotinoid dinotefuran on linden trees in the USA, which
resulted in a significant bumble bee kill (Katchadoorian,
2013), dust generated during planting of a poor-quality
neonicotinoid seed treatment in Germany that affected over
11,000 honey bee colonies (Pistorius et al., 2009), a similar
problem in Italy (APENET, 2011), and dust generation during
planting of neonicotinoid-treated seed in the presence of
seed lubricants in Ontario, Canada (PMRA, 2013; Cutler
et al., 2014b; see http://www.hc-sc.gc.ca/cps-spc/pubs/
pest/_fact-fiche/bee_mortality-mortalite_abeille-eng.php).
It is well established that insecticides can affect individuals
and populations of bees, and the impact will increase with
increased exposure, e.g. if the label does not provide clear
and effective mitigation measures (mitigation selected
for honey bees may not always protect other pollinator
species (Thompson and Hunt, 1999), or the user does
not comply with the label (Johansen, 1977; Kevan et al.,
1990; Thompson and Thorbahn, 2009; Brittain et al., 2010;
Hordzi et al., 2010). However, beyond the small number
of country-level incident schemes there are few data
available on incidents occurring following approved uses
or on the scale of poor practice/non-compliance. There is
evidence of deliberate misuse, i.e., intentional poisoning
(Thompson and Thorbahn, 2009). Albert and Cruz (2006)
present the testimony of owners of an organic farm where
traditional and local knowledge about agricultural practices
were being regenerated in Valencia, Spain. They explained
the problems with a law (called the “pinyolà” decree) that
forbids pollinators in certain areas in this community,
where plantations of clementines (non-native) have been
introduced. Pollination generates seed in clementines,
reducing their market value, therefore pesticides are being
used in order to kill pollinators. There is also evidence that
home and garden pesticide use can impact butterfly and
bumble bee populations (Muratet and Fontaine, 2015).
However, there is also good evidence both from national
incident schemes (Thompson and Thorbahn, 2009) and
from field trials (Gels et al., 2002; Stadler et al., 2003;
Shuler et al., 2005; Larson et al., 2013) that the effects
of insecticides on individuals and populations of honey
bees can be reduced by appropriate mitigation measures,
although the effectiveness of these mitigation measures for
wild bee populations is unclear.
There is limited evidence that increasing the proportion of
natural habitat in the surrounding landscape can buffer the
effects of pesticide use on wild bee abundance and species
richness. For example, Park et al. (2015) observed pesticide
effects on a wild bee community visiting an apple (Malus
domestica) orchard were buffered by increasing proportion
of natural habitat in the surrounding landscape. The direct
consequences for crop yield from pesticide-induced
pollinator losses under field conditions are unresolved
(Kevan et al., 1990; Partap et al., 2001; Richards, 2001).
In the presence of pest pressure, pesticides can enhance
crop yield (Oerke, 2006) but a more limited evidence base
also demonstrates that pesticides used in combination
with managed pollinators can enhance crop yield (Lundin
et al., 2013; Melathopoulos et al., 2014) and environmental
health (Scriber, 2004) and may even improve abundance
of butterflies and bumble bees in urban situations (Muratet
and Fontaine, 2015). More recent reviews have specifically
questioned the widespread use of the neonicotinoid seed
treatments and suggested there is little to no published
evidence to demonstrate economic benefits of these for
farmers (EPA, 2015; Van der Sluijs et al., 2015), although
the number of published trials evaluating this directly is
very small and conflicting data also exist (Afifi et al., 2015;
AgInformatics, 2004). In a recent survey on neonicotinoid
seed treatments (Budge et al., 2015) the benefits of these
seed treatments to crop production in the UK were shown
through reduced applications of other insecticides in autumn
and increased yield in the presence of pest pressure,
although this was variable between years. However, it
also showed an apparent correlation between the scale of
use of imidacloprid as a seed treatment on oilseed rape
seed and increased honey bee colony loss. There was no
apparent correlation with total neonicotinoid use (making
the underlying mechanism of the correlation unclear) and
a number of other factors, such as beekeeping practices
and presence of other forage sources, were not included.
Further large-scale studies are required to develop a greater
understanding of the balance between the benefits of
pesticide use in crop production and the potential risks to
pollinator or other non-target populations.
There have been suggestions that chronic exposure to
certain insecticides (particularly neonicotinoids) may result in
delayed but direct mortality of honey bees (Rondeau et al.,
2014; Sanchez-Bayo and Goka, 2014). However individual
honey bees have been shown to clear imidacloprid rapidly
(Cresswell et al., 2014) and although honey bee colonies
fed high levels of imidacloprid resulted in high adult mortality
and colony failure (Dively et al., 2015), feeding with more
field-realistic exposure levels over an extended period
did not result in increased adult mortality or colony failure
(Faucon et al., 2005; Dively et al., 2015). A similar lack of
adult honey bee mortality following long-term (2-6 weeks)
exposure of colonies has been reported for thiamethoxam
and clothianidin (Pilling et al., 2013; Cutler et al., 2014a;
Sandrock et al., 2014). Recent approaches of using chronic
toxicity (LC 50
) data to assess cumulative toxicity may directly
address such concerns for a wider range of pesticides
(EFSA, 2013).
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2. DRIVERS OF CHANGE OF POLLINATORS,
POLLINATION NETWORKS AND POLLINATION