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Impact of introduced honey bees on native pollination interactions of the endemic Echium wildpretii (Boraginaceae) on Tenerife, Canary Islands


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The aim of this study was to investigate effects of introduced honey bees (Apis mellifera) on native pollination interactions of Echium wildpretii ssp. wildpretii in the sub-alpine desert of Tenerife. We selected two study populations, one dominated by honey bees, while the other was visited by many native insects. During peak activity period of insects, nectar was nearly completely depleted in flowers of the first, but not the latter population. Thus, a high abundance of honey bees may have suppressed visitation by native animals due to exploitative competition. Honey bees stayed longer and visited more flowers on the same inflorescence than native bees, thus potentially promoting self-pollination of the plants. Level of seed set and viability was similar in the two study populations. However, we cannot rule out long-term changes in genetic population structure due to changes in gene-flow patterns caused by foraging behaviour of honey bees vs. native flower-visitors.
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Impact of introduced honey bees on native pollination interactions
of the endemic Echium wildpretii (Boraginaceae) on Tenerife,
Canary Islands
Yoko L. Dupont
, Dennis M. Hansen, Alfredo Valido, Jens M. Olesen
Department of Ecology and Genetics, University of Aarhus, Ny Munkegade Building 540, Aarhus C 8000, Denmark
Received 11 June 2003; received in revised form 6 September 2003; accepted 20 September 2003
The aim of this study was to investigate effects of introduced honey bees (Apis mellifera) on native pollination interactions of
Echium wildpretii ssp. wildpretii in the sub-alpine desert of Tenerife. We selected two study populations, one dominated by honey
bees, while the other was visited by many native insects. During peak activity period of insects, nectar was nearly completely de-
pleted in flowers of the first, but not the latter population. Thus, a high abundance of honey bees may have suppressed visitation by
native animals due to exploitative competition. Honey bees stayed longer and visited more flowers on the same inflorescence than
native bees, thus potentially promoting self-pollination of the plants. Level of seed set and viability was similar in the two study
populations. However, we cannot rule out long-term changes in genetic population structure due to changes in gene-flow patterns
caused by foraging behaviour of honey bees vs. native flower-visitors.
Ó2003 Elsevier Ltd. All rights reserved.
Keywords: Disruption of native mutualisms; Interspecific competition; Apis mellifera; Conservation
1. Introduction
In recent years the impact of introduced honey bees
(Apis mellifera L.) on native flora and fauna has been
much debated. Results of some studies indicate that
foraging patterns and abundance of native pollinators
are altered in the presence of honey bees (Roubik, 1978;
Schaffer et al., 1983; Sugden and Pyke, 1991; Paton,
1993; Wenner and Thorp, 1994; Vaughton, 1996; Gross
and Mackay, 1998; Gross, 2001; Hansen et al., 2002).
Although stressed as important by most researchers, it
has been difficult to investigate potential detrimental
effects of introduced honey bees on food storage or on
reproduction of native bee species (Roubik, 1983; Sug-
den et al., 1996; Butz Huryn, 1997; Steffan-Dewenter
and Tscharntke, 2000; Thorp et al., 2000). The impact of
honey bees on the pollination of native flora includes
effects on pollen dispersal and thus patterns of seed set
and genetic structure of plant populations. Honey bees
are often found to be less efficient pollinators compared
to native flower-visiting animals (Schaffer et al., 1983;
Taylor and Whelan, 1988; Westerkamp, 1991; Paton,
1993; Vaughton, 1996; Gross and Mackay, 1998; Han-
sen et al., 2002). However, other studies have found that
A. mellifera does not adversely affect plant reproductive
success, perhaps due to the numerical abundance of
honey bees compared to native bees (Vaughton, 1992;
Gross, 2001). Furthermore, effects of introduced honey
bees on native flora or fauna are often difficult to assess
due to a lack of suitable control sites (i.e., absence of A.
mellifera). Lastly, patterns induced by honey bees may
be swamped by demographic, stochastic, and environ-
mental variation.
Two features of island pollination networks leave
them susceptible to invasion by introduced generalist
species, such as honey bees: Low species diversity
(Kennedy et al., 2002) and the generalised nature of
interactions (Olesen et al., 2002). Several studies of is-
land ecosystems have reported a decline in both native
bee and bird species visiting flowers in the presence of A.
mellifera (Roubik, 1978; Kato, 1992; Wenner and
Biological Conservation 118 (2004) 301–311
Corresponding author. Fax: +45-86-127-191.
E-mail address: (Y.L. Dupont).
0006-3207/$ - see front matter Ó2003 Elsevier Ltd. All rights reserved.
Thorp, 1994; Kato et al., 1999; Hansen et al., 2002).
Honey bees are widespread in the Canary Islands, and
bee-keeping has been practiced for centuries (M
2000). The colonisation of the islands by A. mellifera is
ancient, and honey bees are considered native based on
mitochondrial DNA data (De la R
ua et al., 1998).
However, A. mellifera is absent from the two eastern
arid islands, Fuerteventura and Lanzarote, where the
climate is dry and the flowering season too short to
support perennial colonies of bees (De la R
ua et al.,
2001). Similar climatic conditions prevail in the semi-
arid, sub-alpine desert zone, found in the mountain re-
gions above 2000 m a.s.l. on Tenerife. Combined with
the geographical isolation provided by both crater rim
and the surrounding pine forest, this suggests a natural
absence of A. mellifera in this habitat. Few, but very
distinct plant species inhabit these altitudes, exposed to
high irradiation, drought, strong winds and extreme
temperature ranges. Most of the native sub-alpine biota
is confined within Teide National Park (18,900 ha),
which is legally protected. However, apiculture is per-
mitted. Every year bee-keepers bring thousands of bee-
hives to the mountain areas of Tenerife during the short
sub-alpine summer (Anon., 2000).
Observations in 2000 and 2001 showed a marked
seasonal shift in the flower-visiting fauna of Echium
wildpretii ssp. wildpretii Pearson ex Hook. f. (Boragin-
aceae) in Teide National Park. In early season, native
passerine birds (Phylloscopus collybita (Vieillot) and
Serinus canarius L.) and native insects visited the red,
nectar-rich flowers. However, coinciding with a sudden
increase in honey bee activity, the birds stopped forag-
ing for nectar (Valido et al., 2002; unpubl. data from
2000). These observations indicate that the introduced
honey bees may alter native pollination interactions.
While the cessation of bird visits was easily observed,
effects on native insects were less obvious. The main
objective of this study is to investigate the relative im-
portance of native insects versus introduced honey bees
as flower-visitors and pollinators of E. wildpretii. Na-
tional Park staff controls the placement and numbers of
beehives, and we were thus able to obtain information
about temporal and geographical distribution of hives
throughout the flowering season.
The controlled framework of bee-keeping in Teide
National Park provides a man-induced ecological ex-
perimental setup, allowing us to assess the impact of
introduced honey bees on reproductive success, and
hence long-term persistence of E. wildpretii. We chose
two study populations, one population close to man-
aged beehives within the caldera and an adjacent pop-
ulation on the outside of the crater rim. We hypothesise
that honey bees, when present in large numbers, deplete
flowers of nectar, thus leading to exploitative competi-
tion with native flower-visitors. Moreover, differences in
foraging patterns of honey bees and natives could alter
pollen flow, and thus affect seed set or seed viability of
E. wildpretii. Therefore, we studied the following key
aspects of the pollination interactions in the two study
populations: (1) Diurnal and seasonal patterns of visi-
tation by native and introduced flower-visitors. (2)
Nectar secretion in animal-excluded flowers and flowers
open to visitation. (3) Patterns of seed set.
2. Materials and methods
2.1. Study system
The study was carried out in Teide National Park,
which is part of the region Las Ca~
nadas, a volcanic al-
luvial plain delimited by a crater rim. The National Park
(NP) covers an area of 18,900 ha and encompasses the
high-altitude sub-alpine zone of Tenerife, Canary Is-
lands. Vegetation is a low and sparse shrub dominated
by a few species.
The Teide bugloss, Echium wildpretii ssp. wildpretii
(Boraginaceae) (E. wildpretii hereafter) was the focal
species of our study. This endemic plant is almost ex-
clusively confined to the sub-alpine mountain zone of
Tenerife (Bramwell and Bramwell, 1990). Although this
subspecies is categorised as rare (G
omez Campo, 1996),
it is locally abundant and the large red-flowered inflo-
rescences are conspicuous in the landscape during the
flowering season. The plant is monocarpic and grows as
a rosette for 5–10 years before producing a single
flowering shoot. The columnar inflorescence is ca. 0.5–
2.5 m in height, a basal diameter of 10–70 cm, tapering
towards the top. Flowers are borne in cymes, which are
arranged spirally on the inflorescence. Cymes have a
total of four (in apical cymes) to 30 flowers (in basal
cymes) with 1–3 open flowers at a time. The flower is
protandrous and is open for 2.5–3 days (Olesen, 1988).
Each flower contains four ovules, developing into a
maximum of four nutlets. Individual plants flower for 3–
5 weeks, but since the time of flowering varies among
individuals, the total flowering period of a population is
ca. 1–1.5 months. In addition to the passerine birds and
honey bees, 16 native insect species and juveniles of the
endemic lacertid lizard Gallotia galloti Dum
eril &
Bibron have been observed visiting the flowers (Valido
et al., 2002).
2.2. Experimental design
Beehives are brought to the park during the flowering
season of the nectar-rich plant species, mainly Descu-
rainia bourgeauana Webb ex O.E. Schulz (Brassicaceae),
Spartocytisus supranubius (L.) Webb & Berth (Fabaceae)
and E. wildpretii. The timing of flowering varies between
years and is tracked by the bee-keepers. Generally, a
maximum of 2500 hives at 25 sites are allowed in the NP
302 Y.L. Dupont et al. / Biological Conservation 118 (2004) 301–311
each year. Each hive houses 10,000–70,000 bees. How-
ever, due to drought and hence scarcity of floral re-
sources, the actual number of beehives placed in the NP
was lower than the allowed maximum in 2001, the year
of our experiment. A total of 1393 beehives were placed
at 17 sites in Teide NP from April 27 to June 10 2001.
Beehives were removed again from May 25 and on-
wards. By July 12th 2001 all hives had been removed
(Fig. 1). Information about numbers of beehives placed
and removed at each site throughout the flowering sea-
son was obtained from Teide NP office. Due to the
limited area of the national park, all populations of E.
wildpretii inside the caldera are located few kilometres
from beehives, and thus visited heavily by honey bees.
Only E. wildpretii populations isolated from bee hives by
a geographical barrier were visited by fewer honey bees.
For this reason, we selected two study populations ca.
2 km apart, separated by the crater rim surrounding Las
Population 1 (hereafter Pop1) was located at ÔCe-
menterio de los TajinastesÕ(28°130N, 16°380W, 2050 m
a.s.l.), on the inner slope of the crater rim. Within 5 km
of Pop1, there were a total of 356 beehives in five lo-
cations. Pop1 was expected to be heavily exploited by
honey bees, as 90% of foraging trips by honey bees are
within a 5-km radius of the colony (Visscher and Seeley,
1982). Vegetation at the study site was dominated by S.
supranubius. Other abundant species were Scrophularia
glabrata Aiton (Scrophulariaceae), Pimpinella cumbrae
Buch ex DC (Apiaceae), Erysimum scoparium (Brouss.)
Wettst. (Brassicaceae), Tolpis webbii Sch. Bip. (Astera-
ceae), and E. wildpretii. Approximately 120 flowering
individuals of E. wildpretii were found at this site.
Population 2 (hereafter Pop2) was located at ÔValle de
los TajinastesÕ(28°120N; 16°370W, 2150 m a.s.l.), less
than 2 km from Pop1, but on the outer side of the crater
rim. This study site was isolated from nearby beehives in
the NP by a 400-m high mountain ridge. Moreover,
Pop2 is isolated from beehives at lower altitudes by pine
forest, a uniform vegetation type consisting of few spe-
cies. We have no information about bee-keeping activity
outside the NP in the nearby pine forest. However, feral
colonies are unlikely to persist in this zone due to the
arid conditions. Thus, Pop2 was predicted to have a low
visitation by A. mellifera. Vegetation was dominated by
S. supranubius and Carlina xeranthemoides L. fil (As-
teraceae). Due to severe drought only very few plant
individuals set flowers in 2001, and most species were
only observed in vegetative condition. E. wildpretii,
represented by 60–70 flowering individuals, was the
most conspicuously flowering species and the only
species offering large amounts of floral resources to
To compare plant characteristics of Pop1 and Pop2,
we measured height of flowering plants, size of inflo-
rescences (length, basal diameter and surface approxi-
mated to a cone), flower density, number of flowers per
cyme and total number of flowers per inflorescence.
Since number of flowers per cyme varied along the
length of the inflorescence, 20 cymes from each of the
lower, mid and upper part of an inflorescence were used
to calculate the average number of flowers per cyme.
Furthermore, we measured the distance to the three
nearest neighbouring conspecific flowering plants and
the size of these. Data from the two study populations
were compared by t-tests when assumptions of para-
metric tests were met. Otherwise the non-parametric
Mann–Whitney U-test was used.
2.3. Flower visitation
To investigate diurnal and seasonal patterns of visi-
tation, we measured visitation rate of native insects and
honey bees from sunrise (ca. 8:00 h) to sunset (ca. 21:00
h) from early to late in the flowering season. We ob-
served bird visits with a binocular from a hideout >10 m
away, this is treated in further detail in another paper,
Valido et al. (2002). In Pop1 insect visitation rates were
recorded on nine observation days (1–27 May) and in
Pop2 on five observation days (15–30 May) interspersed
regularly through the flowering season. To capture
variation due to differences in floral display, spatial lo-
cation and stage of flowering between individual plants,
visitation rate was recorded alternately in 20 study
plants in Pop1 and in six plants in Pop2. Each plant was
observed for 10-min periods consisting of five 2-min
intervals. Depending on the level of visitation, one side
of the whole inflorescence, a proportion (1/2, 1/3 or 1/4)
or a square of 10 10 cm was observed. At the begin-
ning of each 2-min census period, all insect individuals
present on the observed part of the inflorescence were
Fig. 1. Columns indicate total number of beehives within the area of
Teide National Park during 10-day periods (days 1–10, 11–20, etc.).
Number of hives within 5 km of Pop1 is shown in black. Day 1 cor-
responds to April 27, 2001. The last beehives were removed on day 77 –
July 12, 2001.
Y.L. Dupont et al. / Biological Conservation 118 (2004) 301–311 303
counted, and during the 2 min all newcomers were ad-
ded to this number. One visit was defined as an indi-
vidual landing on the inflorescence foraging for nectar
and/or pollen until its departure from the inflorescence.
Since insect individuals were not marked, repeated visits
by the same individual were counted as new visits. Thus,
this census method could lead to an overestimate of
visitation rate especially in the 10 10 cm squares, since
an insect individual may arrive and depart from the
square several times during a single visit to the whole
inflorescence. For this reason we made simultaneous
measurements by two observers, recording visitation
rate in a whole (or half) inflorescence and in 10 10 cm
squares in 32 2-min periods. Based on these data we
constructed regression equations, which were subse-
quently used to convert visitation rate measured in
10 10 squares to visitation rate of the whole inflores-
cence. Visitation rates recorded in 1/2, 1/3 or 1/4 of an
inflorescence were multiplied by 2, 3 and 4, respectively,
to estimate visitation rate per plant. When insects were
counted only in a proportion of the inflorescence, we
alternately observed lower, mid and upper parts of the
inflorescences. Spatial differences in visitation rate were
investigated using seven different plants in Pop2, for
which visits to the upper and lower halves of the inflo-
rescence were recorded. During all visitation observa-
tion periods, we monitored air temperature at the
surface of the inflorescence at mid-inflorescence height.
To investigate differences in foraging behaviour be-
tween honey bees and native insects, we observed the
behaviour of individual honey bees, Anthophora alluaudi
Perez and Eucera gracilipes Perez (Apidae). These two
endemic anthophorid type bees were the most abundant
native visitors. For each bee individual we recorded the
time spent and the number of flowers probed per visit to
an inflorescence.
2.4. Nectar
To investigate diurnal patterns of nectar secretion
and the influence of flower-visitors on nectar availability
in E. wildpretii, we measured nectar volumes using mi-
cro-capillary tubes and sugar concentration using a
hand-held Bellingham–Stanley refractometer. Sampling
was done during peak flowering season, when the ac-
tivity of flower-visiting animals was high (May 13–14 in
Pop1, 15 and 17 May in Pop2). Two adjacent plants of
approximately equal size and stage of flowering were
selected in each study population. The inflorescence of
one plant was covered with fine nylon mesh prior to the
activity period of flower-visitors to exclude animals
(‘‘excluded’’). Flowers of the other plant were left non-
manipulated to be visited by flower-visiting animals
(‘‘open’’). Nectar volume per flower and sugar concen-
tration were measured from sunrise (8:00 h) to sunset
(21:00 h). Sugar weight in sucrose-equivalents was cal-
culated as volume concentration. In each sampling
period, nectar characteristics were measured in 10
flowers of each treatment, five in male and five in female
phase. The light regime (sun or shade) of the sampled
flowers was also recorded. When possible, equal num-
bers of flowers in sun and shade were sampled in each
period. Flowers were removed from the inflorescence
after sampling to avoid repeated sampling. An estimate
of the total lifetime nectar amount offered by an E.
wildpretii individual was calculated as total nectar pro-
duction per plant, based on number of flowers per plant
(using plants of average size and maximum size, re-
spectively), average nectar volume per flower per day
and an average nectar production period of 2 days per
2.5. Seed set and viability
After flowering had ended and fruit development
started, the level of seed set was recorded in 22 plants in
Pop1 and 12 plants in Pop2. In each plant, the infruct-
escence was divided into three parts (lower, mid and
upper). In each part, seed set per flower was counted in
flowers of 20 randomly selected cymes. Seed set was
calculated as a percentage of maximum seed set, which is
four seeds per flower. To assess the importance of ani-
mal pollinators for seed set, we excluded animal visitors
by bagging whole inflorescences of six plants in Pop1
and four in Pop2 before flowering. Using metal wire and
fine nylon mesh, we constructed cages around each in-
florescence, to avoid the mesh from touching (and thus
pollinating) the flowers. At the end of the flowering
season, cages and bags were removed and seed set re-
corded in 60 randomly selected cymes per plant.
Mature seeds from 19 plants of Pop1 and 14 plants of
Pop2 were collected and tested for viability using a 2,3,5
triphenyl tetrazolium chloride (TTC) enzyme activity
test (Heydecker, 1965, 1968). For each plant, 30 seeds
were collected from the lower, mid and upper parts of the
infructescence. Seeds were cut into halves, placed in petri
dishes with the 2,3,5 TTC solution and left in darkness
for 24 h. Seeds were considered viable if the embryo
turned red, a reaction indicating enzyme activity.
2.6. Data analysis
Average visitation rate per 2-min period was calcu-
lated for each 10-min census period for both honey bees
and native insects. Diurnal patterns were analysed by
dividing visitation rate observations into 2-h periods
from 7:00 to 21:00 h, pooling data from all observation
To investigate diurnal changes in nectar availability
of open and excluded flowers, the mean volume and
concentration in ten flowers sampled at each time
were used to represent nectar characteristics (volume,
304 Y.L. Dupont et al. / Biological Conservation 118 (2004) 301–311
concentration and sugar weight) at a given time of the
day. All analyses were run separately for Pop1 and
To analyse patterns of variation in final seed set (log
transformed) at different levels, we carried out an AN-
OVA (GLM procedure) with the levels: populations,
plants within populations, infructescence parts (lower,
mid and upper) within plants and cymes within parts.
All these variables were included in the model as ran-
dom effects and the Type III Sum of Squares were cal-
culated (Shaw and Mitchell-Olds, 1993).
3. Results
3.1. Plant characteristics
Although plants of Pop1 and Pop2 did not differ in
total height, inflorescences of plants in Pop2 were sig-
nificantly shorter, and thus had a smaller surface area.
Moreover, flower density was lower in Pop2, and plants
of this population generally had fewer flowers per in-
florescence per day compared to Pop1, although the
difference was not significant. Furthermore, number of
flowers per cyme was significantly lower in Pop2 (Table
1). Hence, in conclusion, plants in Pop2 had a lower
number of flowers, both on a per day and on a seasonal
basis. Spatial patterns of plants (nearest neighbours) did
not differ between Pop1 and Pop2, although neigh-
bouring plants in Pop2 were smaller, reflecting the
general difference in plant size between populations
(Table 1).
3.2. Flower visitation
A total of 16 native insect species and the introduced
A. mellifera were observed visiting E. wildpretii (see
Table 1 in Valido et al., 2002). Of these, the most
dominant species were A. mellifera,A. alluaudi and E.
gracilipes (Apidae). Diurnal visitation patterns of both
honey bees and native insects were strongly influenced
by temperature in both study populations: Visitation
started at sunrise (when temperatures were 10–11 °C)
and increased steadily until ca. 11:00 h, when the air
temperature levelled off at ca. 23–24 °C (Figs. 2 and 3).
Insect activity decreased again around 18:00 h, when the
temperature decreased. A marked difference in compo-
sition of the flower-visitor pool was found between Pop1
and Pop2. In Pop1, no significant difference was found
in levels of visitation by honey bees and native insects in
the early season (before the 8th of May). However, in
mid and late season (after the 8th of May), visitation by
honey bees increased suddenly, significantly exceeding
visitation by native insects (Fig. 4). This increase was
coincident with an increase in numbers of beehives
placed within the National Park, including the sites close
to Pop1 (Fig. 1). In contrast, in Pop2, level of visitation
by native insects was higher than that of honey bees
both diurnally (Fig. 3) and throughout the season (al-
though only significantly so in mid season) (Fig. 4).
Honey bees, A. alluaudi and E. gracilipes, differed in
foraging patterns, both in time spent per inflorescence
(ANOVA: N¼95, F¼31:31;P<0:0001, R2¼0:40)
and number of flowers probed per visit (ANOVA:
N¼95, F¼21:59, P<0:0001, R2¼0:32) (Table 2).
Table 1
Plant characteristics for E. wildpretii at the two study sites
Characteristic Pop1 Pop2 Statisticsc
Total plant height (cm) 188.5 34.9 (55) 182.5 34.6 (22) t¼0:61n:s:
Height of inflorescence (cm) 139.6 33.8 (55) 111.4 26.0 (25) t¼3:84
Basal diameter of inflorescence (cm) 28.8 11.3 (55) 27.6 6.7 (25) t¼1:10n:s:
Surface of inflorescence (cm2) 6795.3 4236.6 (55) 5087.9 2424.9 (25) t¼3:05
Density of flowers (no. of flowers/cm2) 0.40 0.09 (54) 0.32 0.08 (10) t¼2:89
Open flowers per plant per day 3301 2241 (20) 1836 762 (10) t¼1:97n:s:
Total number of flowers per cymea17.3 4.7 (1320) 15.5 5.5 (660) U¼7:21
Open flowers per cyme per day 1.76 0.52 (250)
Distance to nearest neighbour (m) 6.0 8.1 (22) 13.8 13.6 (12) U¼1:27n:s:
Mean distance to three nearest neighbours (m) 9.1 7.6 (22) 16.5 13.9 (12) U¼1:32n:s:
Surface of nearest neighbour (cm2) 4750.9 3877.9 (22) 5942.6 2370.1 (12) t¼2:43
Nectar volume per flower per day (ll)b9.79 4.61 (111) 4.99 2.13 (80) U¼7:95
Male phase flowers 8.73 3.38 (55) 4.88 1.40 (40) t¼7:35
Female phase flowers 10.82 4.77 (56) 5.09 2.69 (40) U¼5:91
Nectar sugar concentration (%) 15.5 4.1 (111) 15.3 3.3 (76) t¼0:09n:s:
Male phase flowers 13.4 1.9 (55) 14.0 3.2 (36) t¼0:89n:s:
Female phase flowers 16.7 4.6 (56) 16.6 2.9 (40) U¼1:22n:s:
All values are given as means SD (N).
Mean of 20 cymes from each of the upper, middle and lower parts of the inflorescence.
Excluded flowers, based on data from May 13 to 14 (Pop1) and May 15 and 17 (Pop2).
Pop1 and Pop2 were compared using t-tests when data conformed to assumptions of parametric tests, and the non-parametric Mann–Whitney U-
test otherwise. Significance level: ,P<0:005,  ,P<0:0005, n.s., P>0:05.
Y.L. Dupont et al. / Biological Conservation 118 (2004) 301–311 305
Pair-wise comparisons revealed that honey bees spent
significantly more time and visited significantly more
flowers per inflorescence than the native bees. However,
A. alluaudi and E. gracilipes did not differ in foraging
pattern (Tukey–Kramer test). Furthermore, native in-
sects visited the upper part of the inflorescence much
more frequently than the lower part (Wilcoxon signed
ranks test, visitation rate of upper versus lower:
U¼141:5;P<0:001, N¼28). In contrast, visitation
rate of honey bees was higher in the lower part than in
the upper part of an inflorescence (t-test: t¼2:31,
P¼0:029, N¼28).
Two native species of passerine birds, the Common
chiff-chaff (P. collybita) and the Canary (S. canarius),
were commonly observed visiting the flowers for nectar
in early season in both populations. On one occasion we
observed a pair of another passerine, the Blue tit (Parus
caerulius), visiting six inflorescences for nectar in Pop2.
Bird visits continued occasionally in Pop2 throughout
the flowering season. However, after May 8 no bird
visitors were observed in Pop1. The disappearance of
birds was coincident with the increase in honey bee ac-
tivity (Fig. 4).
3.3. Nectar
In Pop1, nectar volume of excluded flowers remained
at a constant level throughout the day (regression:
6 10141822
6 10 14 18 22
6 10141822
6 10141822
Fig. 2. Diurnal variation in visitation rates (individuals/2 min/inflorescence) of native insects and honey bees in Pop1 in early (1–8 May) and late
season (8–27 May). Notice differences in levels of visitation between early and late season (different scales of the Y-axis).
6 10 14 18 22
6 10141822
Fig. 3. Diurnal variation in visitation rates (individuals/2 min/inflorescence) of native insects and honey bees in Pop2.
306 Y.L. Dupont et al. / Biological Conservation 118 (2004) 301–311
lnðvolÞ¼2:00–0.25 h, F¼0:03, P¼0:87, R2¼0:002).
In contrast, nectar volume of open flowers decreased
significantly during the day (regression: lnðvolÞ¼3:99–
0.26 h, F¼7:67, P¼0:01, R2¼0:34). Rapid decrease
to near-zero level occurred from sunrise to ca. 12:00 h,
after which nectar volume remained at a constant low
level until ca. 20:00 h, where a slight increase was ob-
served (Fig. 5). Hence, sugar weight of open flowers was
significantly lower than the level found in excluded
flowers from 09:50 h onwards (Mann–Whitney U-tests,
all P<0:05). In Pop2, nectar volume of excluded
flowers tended to increase slightly during the day (re-
gression: vol ¼)0.06 + 0.46 h, F¼5:61, P¼0:06,
R2¼0:48), albeit only to about half the level of that
found in Pop1 (Table 1). In open flowers, nectar volume
was constant throughout the day (regression: vol ¼3.65
– 0.07 h, F¼0:15, P¼0:71, R2¼0:024) (Fig. 5).
Overall, average values in the four groups (excluded and
open in Pop1 and Pop2, respectively) illustrate the ef-
fects of drought and nectar exploitation by visitors:
Differences in nectar level between excluded flowers of
Pop1 and Pop2 can be explained by extreme drought in
Pop2, while differences between open flowers reflect ex-
ploitation by a flower-visitor fauna dominated by honey
bees (Pop1) versus native insects (Pop2) (Fig. 5, dashed
Nectar secretion in excluded flowers in Pop1 was
significantly influenced by their sexual phase: nectar of
flowers in female phase had a higher sugar concentra-
TIME (h)
6 10141822
TIME (h)
TIME (h)
6 10141822
TIME (h)
Fig. 5. Diurnal variation in nectar volume of animal-excluded and open (non-excluded) flowers of E. wildpretii. Each point represents the mean of 10
flowers. Dashed lines indicate the mean daily levels of nectar in the four groups, plants with excluded and open flowers in Pop1 and Pop2, re-
spectively. Mean of open flowers in Pop1 was calculated excluding data before 10:00 h, when visitation was low.
*n.s. *** ***
1 6 11 16 21 26 31
** *
n.s. n.s.
1 6 11 16 21 26 31
Apis mellifera
Native insects
Fig. 4. Seasonal visitation pattern. Daily averages + SD of visitation
rates of honey bees and native insects in (a) Pop1 and (b) Pop2. Wil-
coxon signed ranks tests were used for comparisons. Significant dif-
ferences are indicated (* for P<0:05, n.s. for non-significance).
Numbers on the X-axis indicate dates in May 2001.
Table 2
Foraging patterns of two native bee species and the introduced honey
bee visiting Echium wildpretii
Species NSeconds/visit Flowers/visit
Eucera gracilipes 62 16.5 23.8 6.7 8.2
Anthophora alluaudi 13 18.7 10.8 9.9 9.5
Apis mellifera 20 137.1131.7 34.5 29.6
Values are given as means SD.
Y.L. Dupont et al. / Biological Conservation 118 (2004) 301–311 307
tion, larger volume, and hence higher sugar weight than
flowers in male phase (Mann–Whitney U-tests, all
P<0:0001) (Table 1). Sugar weight decreased during
the course of the day in male phase flowers and in-
creased steadily in female phase flowers (Fig. 6). In
Pop2, female phase flowers also had a larger sugar
concentration (U¼3:55, P<0:001, N¼76) and sugar
weight (U¼1:99, P<0:05, N¼80) than male phase
flowers. However, differences in nectar volumes were not
significant (U¼1:12, P¼0:26, N¼80). Light regime
(sun or shade) did not affect any nectar characteristics of
flowers in Pop1 or Pop2 (Mann–Whitney U-tests, all
Total lifetime nectar production differed greatly be-
tween populations and among plants within each pop-
ulation. In Pop1, an average-sized plant produced an
estimated amount of 0.54 L during its lifetime, while in
Pop2 this amount was 0.16 L. However, inter-plant
variation was considerable due to variation in size. The
largest plant in Pop1 was estimated to produce 1.75 L,
and in Pop2 the largest lifetime production of a plant
was 0.43 L.
3.4. Seed set and viability
Seed set per flower per plant was slightly higher in
Pop2 than Pop1 (Table 3), although the difference was
not significant (Table 4). Considerable variation in seed
set was found between plant individuals within the
populations, and seed set was positively correlated with
plant height (Pearson r¼0:42, P¼0:01, N¼34).
However, seed set did not differ between populations
when using plant height as a covariate (F¼3:12,
P¼0:11). Furthermore, all the interactions terms cal-
culated in the GLM analysis were highly significant, i.e.,
seed set varied among plants within populations, among
parts of the infructescence within single plants and
among cymes within parts of an infructescence (Table
4). Seed set of infructescences, which had been excluded
from flower-visiting animals, revealed that plants set
some seeds even in the absence of animal pollen vectors
(Table 3). However, seed set was much lower than in the
Table 4
Results of the ANOVA analysis (GLM procedure using Type III Sum of Squares) of seed set (log transformed) at different levels (population, plant
individual, part of the inflorescence and cyme)
Source of variation SS df MS FP
Population 0.49 1 0.49 0.15 0.709
Plant 93.57 20 4.68 1.47 0.263
Part 1.4 2 0.70 6.94 0.002
Cyme 0.72 19 0.04 0.96 0.527
Population Plant 34.14 10 3.41 129.81 <0.001
Plant Part 4.21 40 0.10 4.00 <0.001
Part Cyme 1.53 38 0.04 1.53 0.020
Error 882.47 33,556 0.03
All variables were treated as random effects. The obtained model was statistically significant (F¼2912:8; P¼0:008).
6 10141822
6 10141822
Fig. 6. Diurnal variation in sugar weight of nectar of excluded flowers in male and female phase. Each point represents the mean of five flowers. Data
were only obtained for Pop1.
Table 3
Seed set and viability in open pollinated flowers, and flowers excluded
from flower-visitors
Parameter Pop1 Pop2
Seed setb, open 48.0 27.1 (22,732) 54.2 23.7 (10,955)
Seed setb, excluded 23.3 27.2 (6020) 29.6 30.7 (3744)
Seed viabilityc, open 84.4 9.1 (57) 83.9 10.2 (48)
Means SD (N)a.
Seed viability N, number of viability tests. Seed set N, number of
Percentage of maximum seed set (four seeds per flower).
Percentage viable seeds in viability tests of 30 seeds.
308 Y.L. Dupont et al. / Biological Conservation 118 (2004) 301–311
controls for both Pop1 (U¼58:71, P<0:00005,
N¼28;752) and Pop2 (U¼44:26, P<0:00005,
There were no significant differences in seed viability
between populations (Table 3). Viability differed signif-
icantly among plants in Pop2 (ANOVA: df ¼15,
F¼3:84, P¼0:0007, R2¼0:64), but not in Pop1
(df ¼18, F¼1:68, P¼0:09, R2¼0:44).
4. Discussion
4.1. Impact of honey bees on native insects and birds
The potential impact of introduced A. mellifera in
native systems is a major concern, but few studies pro-
vide quantitative data and clear-cut evidence of the ef-
fects of honey bees (Sugden et al., 1996; Butz Huryn,
1997; Kearns et al., 1998). Results of this study show
that the site dominated by honey bees is characterised by
low visitation rate by native insects throughout the
season. In contrast, a high level of visitation by native
insects is maintained during the flowering season in
Pop2, which had a lower level of honey bee visitation.
The high abundance of honey bees may have resulted in
exploitative competition between honey bees and native
insects, as nectar standing crops were reduced to near-
zero levels in Pop1, but not in Pop2, despite the more
arid conditions prevailing there (Fig. 5). On the other
hand, the level of visitation by native insects remained
constant throughout the flowering season in Pop1, and
thus appeared to be unaffected by the emergence of
honey bees (Fig. 4). Two alternative scenarios may ex-
plain this pattern: (1) abundance of native insects was
limited by low temperatures in early season, and later
suppressed by honey bee dominance or (2) because ex-
tensive honey bee keeping has been practiced for cen-
turies in Las Ca~
nadas (M
endez, 2000), we may be
observing the Ôghost of competition pastÕ, i.e., numbers
of native flower-visitors may already have been reduced
by honey bee dominance in Pop1. It would be interest-
ing to address this question in a future study.
Exploitative competition may also be acting between
honey bees and passerine birds. In two consecutive years
we observed that nectar-feeding birds ceased flower-
visitation in Pop1 when honey bees became abundant.
At this stage it is not possible to determine if this pattern
is connected to depletion of nectar or to a seasonal
change in the diet of birds related to, e.g., availability of
insects or breeding activity. E. wildpretii is one of the
largest nectar resources in the otherwise very dry and
sparsely vegetated environment of Las Ca~
nadas. Thus,
overexploitation of E. wildpretii could force native
flower-visitors to switch to other, less profitable floral
resources. Furthermore, preliminary observations of
flower-visiting insects in a population of E. wildpretii
ssp. trichosiphon on La Palma Island revealed a lower
abundance of A. mellifera, higher levels of standing
nectar crop and a higher diversity of native insects
compared to the honey bee dominated population on
Tenerife (unpubl. data).
As mentioned in Section 1, other studies have shown
that introduced honey bees negatively affect visitation
rate and species diversity of native flower-visitors (Kato,
1992; Wenner, 1993; Kato et al., 1999; Wenner et al.,
2000; Gross, 2001; Hansen et al., 2002). An interesting
question is whether this pattern translates into reduced
reproductive output of native insects, and hence
threatens their long-term persistence in Teide National
Park. In a long-term study, Roubik and Wolda (2001)
found no decreases in population size of native insects in
the presence of africanised honey bees in a Central
American rain forest. On the other hand, an experi-
mental study in an Australian tree-grass plain showed
population declines of the native bee Exoneura asimil-
lima in the presence of managed bee hives, possibly due
to resource competition with honey bees (Sugden and
Pyke, 1991). Resource level has also been shown to af-
fect reproduction in the leaf-cutter bee Megachile api-
calis (Kim, 1999). Obviously, the long-term impact of A.
mellifera on a native pollination system varies between
regions and habitat types, and thus extrapolation of
above results to desert areas like Las Ca~
nadas should be
made with extreme caution (Paton, 1993). More long-
term studies are clearly necessary to assess the impact of
A. mellifera at the population level of native flower-
4.2. Impact of honey bees on E. wildpretii
Introduced honey bees are known to reduce fitness of
some native plant species (Gross and Mackay, 1998).
However, in other cases, seed set has been shown to be
unaffected (Vaughton, 1992) or pollination even aug-
mented by the presence of honey bees (Paton, 1993;
Gross, 2001). In our study, seed set was not significantly
different in the study population dominated by honey
bees and the population visited predominantly by native
insects, the level of seed set being only slightly lower in
the former. Hence, the effect of A. mellifera on seed
production is minor, if any. Some studies have shown
that honey bees are poorer pollinators than native spe-
cies (Westerkamp, 1991; Freitas and Paxton, 1998;
Gross and Mackay, 1998; Hansen et al., 2002), but a
high abundance may compensate for lowered pollina-
tion efficiency (Butz Huryn, 1995; Kraemer and Schmitt,
1997; England et al., 2001). Moreover, breeding system
of the plant influences level of seed set. E. wildpretii was
capable of producing a considerable number of seed
without being visited by animals, and thus may to some
extent be pollinated by wind or even set seeds by au-
togamy. On the other hand, open pollination increases
Y.L. Dupont et al. / Biological Conservation 118 (2004) 301–311 309
seed set, and most aspects of anthesis seem to be related
to animal pollination: Most flowers open in the morn-
ing, live for 2.5 days, with approximately 1 day in male
phase and 1 day in female phase (Olesen, 1988). Our
study showed that flowers had two peaks in sugar
weight of nectar, one in male phase and one in female
phase, separated by a minimum during the sexual
transition phase (Fig. 6). Thus, flowers are most at-
tractive to flower-visitors during peak pollen presenta-
tion and again during stigma receptivity.
It is difficult to assess the influence of animal visita-
tion for variation in seed set, superimposed upon the
background reproductive output by spontaneous au-
togamy and wind pollination. Our analyses indicate a
positive correlation between plant size and seed set,
which is concordant with a preference of animals to visit
larger plants (Valido et al., 2002). However, the same
pattern could be explained by resource allocation in
semelparous organisms, larger plants having accumu-
lated more resources before reproduction.
4.3. Future perspectives
Although level of seed set can be slightly affected by
foraging patterns of the visitors, the primary impact of
introduced honey bees may be changes in pollen flow,
and thus genetic structure of the plant population. In
contrast to native insects and birds, honey bees visited
many flowers on each inflorescence and rarely moved
between plants, which is likely to promote geitonogamy
and hence inbreeding. However, how and if this affects
long-term persistence of the plant population is un-
known. We found no differences in seed viability be-
tween plants pollinated mainly by honey bees or native
insects. One explanation could be that E. wildpretii is
relatively unaffected by inbreeding, since it is capable of
producing a large number of seeds by selfing alone (up
to 50% of the seed set level of open pollinated plants).
Plants in Las Ca~
nadas have been pollinated by honey
bees since these were introduced in the 16th century
endez, 2000). Thus, honey bees have influenced the
pattern of pollen transfer in 80–100 plant generations,
which may have contributed to purging of deleterious
alleles expressed through inbreeding. On the other hand,
inbreeding effects in life stages other than seed viability
cannot be ruled out. In a close relative, Echium vulgare
L., no effect of selfing was found at the stage of seed
production (Mensler et al., 1997). Yet, late-acting in-
breeding depression in male and female function of
offspring derived from selfing has been reported
(Mensler et al., 1999). Future studies should address
patterns of genetic variation in populations of E.
wildpretii pollinated predominantly by introduced honey
bees versus those pollinated by native animals. For in-
stance, in Grevillea macleayana (Proteaceae), which is
visited mainly by native birds and introduced honey
bees, outcrossing rate was reduced significantly when
birds were excluded from the inflorescences (England
et al., 2001). Furthermore, an interesting question is the
role of birds as pollinators and potential long-distance
pollen vectors of E. wildpretii in early flowering season,
before the onset of bee-keeping activities.
Many studies call for further investigation of the ef-
fects of honey bees on the reproductive output of plants
and long-term persistence of native flower-visiting animal
populations. However, impact of honey bees are difficult
to disentangle from confounding biotic and abiotic fac-
tors. Las Ca~
nadas offers a unique and simple study sys-
tem. The total flowering season of the sub-alpine desert
system is short (ca. 2 months) and the plant–flower-visi-
tor network is simple, consisting of a few species isolated
by the crater rim and surrounding pine forest (Dupont
et al., 2003). Furthermore, placement of beehives can be
controlled both spatially (as in Paton, 1993) and tem-
porally, creating gradients of honey bee visitation pres-
sure over the season and between plant populations.
We are grateful to M. Durb
an, A. Ba~
nares and J.
on (Teide National Park, Tenerife), and A. Palo-
mares and A. Revol
e (La Caldera de Taburiente Na-
tional Park, La Palma) for their collaboration, for
information about the bee keeping activities in the Na-
tional Parks, for research permissions and for letting us
stay at the field stations. We also thank C. Skov, E. P
and E. Portellano for help in the field and for fruitful
discussions. A. Sølling carried out the viability tests. The
manuscript was improved through comments and sug-
gestions by C. Rasmussen. This project received finan-
cial support from Mr. and Mrs. FiedlerÕs grant (to Y.L.
Dupont) the Augustinus Foundation (to Y.L. Dupont)
and the Danish National Science Research Council (to
J.M. Olesen). During the writing of this paper A. Valido
was supported by a Marie Curie individual fellowship
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Y.L. Dupont et al. / Biological Conservation 118 (2004) 301–311 311
... Beekeeping activities can produce unnaturally high local densities of honey bees, and therefore, there is a strong potential for competition between them and wild pollinators (Goulson 2003, Geslin et al. 2017. Managed honey bees reduce pollen and nectar availability (Dupont et al. 2004, Torné-Noguera et al. 2016, Cane and Tepedino 2017, competitively displace wild pollinators from floral resources (Dupont et al. 2004, Shavit et al. 2009, Artz et al. 2011, Lindström et al. 2016, Ropars et al. 2019) and influence their foraging behaviour (Thomson 2004, Walther-Hellwig et al. 2006, Artz et al. 2011. Ultimately, the presence of honey bees can reduce the size, biomass and/or reproduction of wild bees (Thomson 2004, Goulson and Sparrow 2009, Elbgami et al. 2014, Torné-Noguera et al. 2016. ...
... Beekeeping activities can produce unnaturally high local densities of honey bees, and therefore, there is a strong potential for competition between them and wild pollinators (Goulson 2003, Geslin et al. 2017. Managed honey bees reduce pollen and nectar availability (Dupont et al. 2004, Torné-Noguera et al. 2016, Cane and Tepedino 2017, competitively displace wild pollinators from floral resources (Dupont et al. 2004, Shavit et al. 2009, Artz et al. 2011, Lindström et al. 2016, Ropars et al. 2019) and influence their foraging behaviour (Thomson 2004, Walther-Hellwig et al. 2006, Artz et al. 2011. Ultimately, the presence of honey bees can reduce the size, biomass and/or reproduction of wild bees (Thomson 2004, Goulson and Sparrow 2009, Elbgami et al. 2014, Torné-Noguera et al. 2016. ...
... In agreement with several previous correlational and experimental studies (Dupont et al. 2004, Shavit et al. 2009, Artz et al. 2011, Goras et al. 2016, Lindström et al. 2016, Herrera 2020, we also detected a decrease in overall wild bee abundance with increasing honey bee visitation rate. However, the effect we detected was relatively weak compared to the effect on abundance of variables related to landscape disturbance. ...
Full-text available
Maintaining the diversity of wild bees is a priority for preserving ecosystem function and promoting stability and productivity of agroecosystems. However, wild bee communities face many threats and beekeeping could be one of them, because honey bees may have a strong potential to outcompete wild pollinators when placed at high densities. Yet, we still know little about how beekeeping intensity affects wild bee diversity and their pollinator interactions. Here, we explore how honey bee density relates to wild bee diversity and the structure of their pollination networks in 41 sites on 13 Cycladic Islands (Greece) with similar landscapes but differing in beekeeping intensity. Our large‐scale study shows that increasing honey bee visitation rate had a negative effect on wild bee species richness and abundance, although the latter effect was relatively weak compared to the effect of other landscape variables. Competition for flowering resources (as indicated by a resource sharing index) increased with the abundance of honey bees, but the effect was more moderate for wild bees in family Apidae than for bees in other families, suggesting a stronger niche segregation in Apidae in response to honey bees. Honey bees also influenced the structure of wild bee pollination networks indirectly, through changes in wild bee richness. Low richness of wild bees in sites with high honey bee abundance resulted in wild bee networks with fewer links and lower linkage density. Our results warn against beekeeping intensification in these islands and similar hotspots of bee diversity, and shed light on how benefits to pollination services of introducing honey bees may be counterbalanced by detriments to wild bees and their ecosystem services.
... Novel mutualisms that benefit endangered or declining species can be seen as beneficial for an ecosystem, but clearly identifying whether a novel honey bee association exists at the expense of native species can be difficult. For example, removing honey bees from islands has been suggested to benefit species of native bees and birds ( Dupont et al., 2004 ;Kaiser-Bunbury et al., 2010 ;Kato et al., 1999 ), but control sites without honey bees are often lacking ( Dupont et al., 2004 ). In addition, extinct or rare species may not be able to recover after local honey bee removals, making these knowledge gaps even more obtuse. ...
... Novel mutualisms that benefit endangered or declining species can be seen as beneficial for an ecosystem, but clearly identifying whether a novel honey bee association exists at the expense of native species can be difficult. For example, removing honey bees from islands has been suggested to benefit species of native bees and birds ( Dupont et al., 2004 ;Kaiser-Bunbury et al., 2010 ;Kato et al., 1999 ), but control sites without honey bees are often lacking ( Dupont et al., 2004 ). In addition, extinct or rare species may not be able to recover after local honey bee removals, making these knowledge gaps even more obtuse. ...
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Worldwide, the use of managed bees for crop pollination and honey production has increased dramatically. Concerns about the pressures of these increases on native ecosystems has resulted in a recent expansion in the literature on this subject. To collate and update current knowledge, we performed a systematic review of the literature on the effects of managed and introduced bees on native ecosystems, focusing on the effects on wild bees. To enable comparison over time, we used the same search terms and focused on the same impacts as earlier reviews. This review covers: (a) interference and resource competition between introduced or managed bees and native bees; (b) effects of introduced or managed bees on pollination of native plants and weeds; and (c) transmission and infectivity of pathogens; and classifies effects into positive, negative, or neutral. Compared to a 2017 review, we found that the number of papers on this issue has increased by 47%. The highest increase was seen in papers on pathogen spill-over, but in the last five years considerable additional information about competition between managed and wild bees has also become available. Records of negative effects have increased from 53% of papers reporting negative effects in 2017 to 66% at present. The majority of these studies investigated effects on visitation and foraging behaviour. While only a few studies experimentally assessed impacts on wild bee reproductive output, 78% of these demonstrated negative effects. Plant composition and pollination was negatively affected in 7% of studies, and 79% of studies on pathogens reported potential negative effects of managed or introduced bees on wild bees. Taken together, the evidence increasingly suggests that managed and introduced bees negatively affect wild bees, and this knowledge should inform actions to prevent further harm to native ecosystems.
... Island insect pollinator communities are largely made up of small, generalist solitary bees and/or flies, with butterflies and social bees being far less common (Abe, 2006), although through human-mediated movement, western honey bee now dominates pollination networks on many islands across the world (e.g. Tenerife, Canary Islands [Dupont et al., 2004] and the Hawaiian archipelago [Valenzuela, 2018]). There are exceptions, however, and many islands are home to unique, specialised plant-pollinator interactions (Abe, 2006), perhaps the most notable example being the island phenomenon of lizard pollination (Olesen & Valido, 2003). ...
... Irrespective of any pathogens, this can, and has had a big effect on ecosystems in itself. For example, honey bees have been shown to outcompete native pollinators in Tasmania (Goulson et al., 2002), the Bonin Islands (Kato et al., 1999) and Tenerife (Dupont et al., 2004). ...
Island ecosystems, which often contain undescribed insects and small populations of single island endemics, are at risk from diverse threats. The spread of pathogens is a major factor affecting not just pollinator species themselves, but also posing significant knock-on effects to often fragile island ecosystems through disruption of pollination networks. Insects are vulnerable to diverse pathogens and these can be introduced to islands in a number of ways, e.g. via the introduction of infected managed pollinator hosts (e.g. honey bees and their viruses, in particular Deformed wing virus), long-range migrants (e.g. monarch butterflies and their protozoan parasite, Ophryocystit elektroscirrha) and invasive species (e.g. social wasps are common invaders and are frequently infected with multi-host viruses such as Kashmir bee virus and Moku virus). Furthermore, these introductions can negatively affect island ecosystems through outcompeting native taxa for resources. As such, the greatest threat to island pollinator communities is not one particular pathogen, but the combination of pathogens and introduced and invasive insects that will likely carry them.
... Of particular note, we found that many native plants were most frequently visited by the introduced honeybee A. mellifera, and that the most common native bee, T. carbonaria (Bilpin only), was rarely seen visiting native plants during the study period, despite showing high abundance on introduced weeds and apple flowers; this could have consequences for native plant reproduction. There is evidence that A. mellifera can differ in its foraging behavior, particularly in regard to the number of intra and inter plant movements between flowers, compared to native pollinator species [18,19,71]. However, it is often unclear how differences in honeybee foraging behavior, in turn, can affect seed set [19,72,73]. ...
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Co-flowering plants can experience an array of interactions, ranging from facilitation to competition, the direction and strength of which are often dependent on the relative abundance and diversity of the plant species involved and the foraging behavior of their pollinators. Understanding interactions between plant–pollinator networks and how they change over time is particularly important within agricultural systems, such as apples, that flower en masse and that also contain non-crop co-flowering species both within the farm and the surrounding landscape. We determined the degree of overlap between pollinator networks on two varieties of apple (Granny Smith and Pink Lady) and co-flowering plant species within orchards and the wider vegetation matrix in two apple-growing regions (Orange and Bilpin) in Australia. We surveyed plant–pollinator interactions at key stages of the cropping cycle: before mass flowering; during king, peak and late blooms; and, finally, once apple flowering had finished. Overall, we found considerable overlap in the flower visitor assemblage on apples and co-flowering species within the orchard. The introduced honeybee (Apis mellifera) was the most frequent flower visitor to all three vegetation types at all times in Orange. However, in Bilpin, both a native stingless bee (Tetragonula carbonaria) and A. mellifera were highly frequent visitors, both on- and off-crop. Numerous native bees, flies and Lepidoptera also commonly visited apple and co-flowering species within orchards in both locations. We found that native-bee and honeybee visitation to apple flowers was positively correlated with co-flowering species richness (within the orchard and the wider matrix); however, visitation by native bees decreased as the area of co-flowering species in the surrounding landscape increased. Our study highlights the importance of maintaining diverse co-flowering plant communities within the local landscape to increase and support a wide variety of pollinators in horticultural production systems.
... However, resources are not always at surplus levels (Zimmerman and Pleasants 1982;Dupont et al. 2004), and honeybees are documented to remove up to 97.2% of the nectar and 99.0% of pollen produced by some native Australian flora (Paton 1990). A study on Tasmanian leatherwood (Eucryphia lucida), a nectar resource highly sought after by commercial honey producers, found that honeybees rapidly removed pollen, and standing crops of nectar sugar were significantly depressed at apiary sites compared with control sites (situated 2 km away from apiaries), where pollen remained in flowers until the female phase (Mallick and Driessen 2009). ...
... This difference in pollination success suggests that native plants are relatively specialized to native bumblebees, and therefore, invasive bumblebees could be poor substitutes for local ones, and that even at medium invasion densities, the fertility of native plants is compromised. Invasive bumblebees also show patterns of pollen transport different to those of native pollinators, potentially altering the genetic structure of plant populations-as it was reported in pollination networks invaded by honeybees [98][99][100]. For raspberry crops in northwestern Patagonia, disproportionate abundance of invasive B. terrestris relative to owers altered the historical cost-bene t balance of pollination interactions and drove this mutualism toward antagonism by decreasing fruit yields [101,102]. ...
Adequate pollination is fundamental to optimize reproduction and yield of most flowering plants, including many staple food crops. Plants depending on insect pollination rely heavily on many wild species of solitary and social bees, and declines or absence of bees often hampers crop productivity, prompting supplementation of pollination services with managed bees. Though honeybees are the most widely deployed managed pollinators, many high-value crops are pollinated more efficiently by bumblebees ( Bombus spp.), prompting domestication and commercial rearing of several species. This led to a blooming international trade that translocated species outside their native range, where they escaped management and invaded the ecosystems around their deployment sites. Here, we briefly review the history of bumblebee invasions and their main impacts on invaded ecosystems, and close by discussing alternatives to the use of commercially reared bumblebees to enhance crop pollination. As evidence of widespread negative effects on local ecosystems of bumblebee invasions builds up, bumblebee trade adds to the list of examples of "biological" strategies devised to solve agricultural problems that ended up being far from the "green," eco-friendly solutions they were expected to be.
... Studies assessing the impact of honeybees on wild bees in agricultural and natural settings found evidence for negative effects of rising honeybee hive densities on wild bee communities and visitation rates through competition for food resources (Torné-Noguera et al. 2016;Dupont et al. 2004;Lindström et al. 2016;Mallinger et al. 2017;Geldmann & González-Varo 2018;Geslin et al. 2017). With urban beekeeping being on the rise, competition from increased numbers of honey bee hives has been identified as one of the main threats for urban pollinator conservation (Baldock 2020). ...
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While urban beekeeping is on the rise, data on the role of wild bee communities as crop pollinators in cities is still scarce. We analysed wild bee visitation rates on apple, plum, cherry, pear, blackberry, raspberry, and strawberry in a Bavarian city with a very high honeybee density of c. 19 hives/km2. During 137.5 hours of observation time, we observed 52 wild bee species on the studied crop plants. During more than 50 h of observation time on fruit trees in flower, we found that wild bees provided 41% of the total bee visits, honeybees the remaining 59%. Honeybee hive density had a significantly negative effect on wild bee abundance. Bumblebees appeared more tolerant to poor weather conditions than all other bee groups. Wild bee species richness on apple flowers was not significantly impacted by flower diversity in the surroundings of the trees. Together, our results suggest that species-rich wild bee communities in urban areas are important for pollination success in common fruit crops, especially under unstable spring weather conditions. Bee-friendly management of urban spaces should be prioritised to support wild bee communities as well as the increasing number of honeybees in cities.
... On the Canary Islands, seed set was compared in two populations of Echium wildpretii, one pollinated by an insect community comprised of mostly the exotic A. mellifera, and the other pollinated by mostly native insect pollinators. The two E. wildpretii populations did not differ in seed set nor seed viability despite higher bee densities at the test site (Dupont et al. 2004). In Australia, the endemic shrub Dillwynia juniperina was not more pollenlimited in sites with more exotic A. mellifera than native insect pollinators compared to sites with mostly natives (Gross 2001). ...
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Insect pollinators have been relocated by humans for millennia and are, thus, among the world’s earliest intentional exotic introductions. The introduction of managed bees for crop pollination services remains, to this day, a common and growing practice worldwide and the number of different bee species that are used commercially is increasing. Being generalists and frequently social, these exotic species have the potential to have a wide range of impacts on native bees and plants. Thus, understanding the consequences of introduced species on native pollinator systems is a priority. We generated a global database and evaluated the impacts of the two main groups of invasive bees, Apis mellifera and Bombus spp., on their pollination services to native flora and impacts on native pollinators. In a meta-analysis, we found that per-visit pollination efficiency of exotic pollinators was, on average, 55% less efficient than native pollinators when visiting flowers of native species. In contrast to per-visit pollination efficiency, our meta-analysis showed that visitation frequency by exotic pollinators was, on average, 80% higher than native pollinators. The higher visitation frequency of exotic pollinators overcame deficiencies in pollen removal and transfer resulting in seed/fruit set levels similar to native pollinators. Also, evidence showed that exotic pollinators can displace native insect and bird pollinators. However, the direct effects of exotic insect pollinators on native pollination systems can be context dependent, ranging from mutualism to antagonism.
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If human population growth is not controlled, natural areas must be sacrificed. An alternative is to create more habitat, terraforming Mars. However, this requires establishment of essential, ecosystem services on a planet currently unamenable to Terran species. Shorter term, assembling Terran-type ecosystems within contained environments is conceivable if mutually supportive species complements are determined. Accepting this, an assemblage of organisms that might form an early, forest environment is proposed, with rationale for its selection. A case is made for developing a contained facsimile, old growth forest on Mars, providing an oasis, proffering vital ecosystem functions (a forest bubble). It would serve as an extraterrestrial nature reserve (ETNR), psychological refuge and utilitarian botanic garden, supporting species of value to colonists for secondary metabolites (vitamins, flavours, perfumes, medicines, colours and mood enhancers). The design presented includes organisms that might tolerate local environmental variance and be assembled into a novel, bioregenerative forest ecosystem. This would differ from Earthly forests due to potential impact of local abiotic parameters on ecosystem functions, but it is argued that biotic support for space travel and colonization requires such developments. Consideration of the necessary species complement of an ETNR supports a view that it is not humanity alone that is reaching out to space, it is life, with all its diverse capabilities for colonization and establishment. Humans cannot, and will not, explore space alone because they did not evolve in isolation, being shaped over aeons by other species. Space will be travelled by a mutually supportive system of Terran organisms amongst which humans fit, exchanging metabolites and products of photosynthesis as they have always done.
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Research in Macaronesia has led to substantial advances in ecology, evolution and conservation biology. We review the scientific developments achieved in this region, and outline promising research avenues enhancing conservation. Some of these discoveries indicate that the Macaronesian flora and fauna is composed of rather young lineages, not Tertiary relicts, predominantly of European origin. Macaronesia also seems to be an important source region for back-colonization of continental fringe regions on both sides of the Atlantic. This group of archipelagos (Azores, Madeira, Selvagens, Canary Islands, and Cabo Verde) has been crucial to learn about the particularities of macroecological patterns and interaction networks on islands, providing evidence for the development of the General Dynamic Model of oceanic island biogeography and subsequent updates. However, in addition to exceptionally high richness of endemic species, Macaronesia is also home to a growing number of threatened species, along with invasive alien plants and animals. Several innovative conservation and management actions are in place to protect its biodiversity from these and other drivers of global change. The Macaronesian Islands are a well-suited field of study for island ecology and evolution research, mostly due to its special geological layout with 40 islands grouped within five archipelagos differing in geological age, climate and isolation. A large amount of data is now available for several groups of organisms on and around many of these islands. However, continued efforts should be made towards compiling new information on their biodiversity, to pursue various fruitful research avenues and develop appropriate conservation management tools.
Like other members of the Canarian flora, Echium wildpretii has often been considered to be pollinated by birds. Although the red flowers of Echium wildpretii secrete relatively large amounts of dilute nectar, birds could not be observed at the species. The main flower visitors are honeybees and the solitary bee Anthophora alluaudi, the latter being the most effective pollinator. Both bee species show an increase in nectar sampling activity during the hot daytime which may be a reaction to higher water demands. The highly concentrated nectar becomes diluted during the course of the day, thus meeting the energetic requirements of Anthophora bees in the morning as well as their water needs later on. Almost all the nectar produced is harvested by the bees, which probably accounts for the absence of bird visits.
Do honey bees compete for food with other bee species in nature? This question has been the focus of considerable scientific and political attention in recent years, especially, but not exclusively in Australia. In this article we provide the background and rationale of the argument and present scientific studies which have attempted to provide evidence. We also suggest some approaches to dealing with real issues related to honey bee competition, bee conservation, and honey bee management.
Ecological studies typically involve comparison of biological responses among a variety of environmental conditions. When the response variables have continuous distributions and the conditions are discrete, whether inherently or by design, then it is appropriate to analyze the data using analysis of variance (ANOVA). When data conform to a complete, balanced design (equal numbers of observations in each experimental treatment), it is straightforward to conduct an ANOVA, particularly with the aid of the numerous statistical computing packages that are available. Interpretation of an ANOVA of balanced data is also unambiguous. Unfortunately, for a variety of reasons, it is rare that a practicing ecologist embarks on an analysis of data that are completely balanced. Regardless of its cause, lack of balance necessitates care in the analysis and interpretation. In this paper, our aim is to provide an overview of the consequences of lack of balance and to give some guidelines to analyzing unbalanced data for models involving fixed effects. Our treatment is necessarily cursory and will not substitute for training available from a sequence of courses in mathematical statistics and linear models. It is intended to introduce the reader to the main issues and to the extensive statistical literature that deals with them.
The red corolla colour, the relatively large volume of nectar and the low concentration of sucrose in the nectar of Echium wildpretii are traits related to bird pollination, while its ultraviolet colour and floral morphology may attract bees. It is suggested that the species may serve as a water resource to solitary desert bees in nature.
Do honey bees compete for food with other bee species in nature? This question has been the focus of considerable scientific and political attention in recent years, especially, but not exclusively in Australia. In this article we provide the background and rationale of the argument and present scientific studies which have attempted to provide evidence. We also suggest some approaches to dealing with real issues related to honey bee competition, bee conservation, and honey bee management.
Previous studies of introduced honey bees foraging at Agave schottii flowers suggest that Apis mellifera preferentially exploits the most productive patches of flowers and thereby reduces the standing crop of available nectar and the utilization of these sites by native bees. Results of experiments undertaken to evaluate this hypothesis are given and discussed using Apis, Bombus and Xylocopa. -Authors
We present data on late-acting inbreeding depression in pollen performance, siring success and seed production in Echium vulgare. Pollen viability and rate of pollen-tube growth were both lower for pollen from plants derived from selfing than for pollen from plants derived from outcrossing. Pollen tube numbers within the styles did not differ for pollen from plants derived from selfing or outcrossing. A pollination experiment with two mixtures of pollen from plants derived from selfing or outcrossing, revealed a significant decline of 55% in siring success for pollen from plants derived from selfing. A second experiment with a complete diallel design revealed inbreeding depression for both siring success of the offspring (32.8%) and a decline in seed production of the offspring (34.8%–40.6%). In addition, results indicated a heritable component for seed number per flower. Offspring fitness, measured as seed production and siring ability, can be severely affected by late-acting inbreeding depression. Inbreeding depression values for male and female functions were not correlated. Both functions must therefore be considered when calculating inbreeding depression.