Invasive bees and their impact
Marcelo A. Aizen
*, Marina P. Arbetman
, Natacha P. Chacoff
Vanina R. Chalcoff
, Peter Feinsinger
, Lucas A. Garibaldi
Lawrence D. Harder
, Carolina L. Morales
, Agustín Sáez
Adam J. Vanbergen
Instituto de Investigaciones en Biodiversidad y Medioambiente (INIBIOMA), Universidad Nacional del
Comahue-CONICET, San Carlos de Bariloche, Rio Negro, Argentina
Instituto de Ecologı
´a Regional (IER), CONICET—Universidad Nacional de Tucuma
´n, San Miguel de
Facultad de Ciencias Naturales e Instituto Miguel Lillo, Universidad Nacional de Tucuma
´n, San Miguel de
Centro de Estudio y Aplicacio
´n del Ciclo de Indagacio
´n, Facultad de Ciencias Naturales e Instituto Miguel
Lillo, Universidad Nacional de Tucuma
´n, San Miguel de Tucuma
Universidad Nacional de Rı
´o Negro, Instituto de Investigaciones en Recursos Naturales,
´a y Desarrollo Rural, San Carlos de Bariloche, Rı
´o Negro, Argentina
Consejo Nacional de Investigaciones Cientı
´ficas y Tecnicas, Instituto de Investigaciones en Recursos
´a y Desarrollo Rural, San Carlos de Bariloche, Rı
´o Negro, Argentina
Department of Biological Sciences, University of Calgary, Calgary, AB, Canada
Agroecologie, AgroSup Dijon, INRAE, Universite de Bourgogne Franche-Comte, Dijon, France
*Corresponding author: e-mail address: firstname.lastname@example.org
1. Introduction 50
2. Plant-pollinator interaction as a benefit-cost relation 53
2.1 Direct costs 54
2.2 Indirect costs 57
3. Invasive bees 59
3.1 The western honey bee 60
3.2 Bumble bees 61
3.3 Other bees 62
4. Drivers of bee invasion success 62
4.1 Bee introductions and trade 63
4.2 Ecological opportunities and displacement of native bees 65
4.3 Anthropogenic disturbance 67
5. Consequences of bee invasions for crop pollination 68
5.1 Interaction of alien bees with crop flowers 69
5.2 Visit frequencies of alien bees 71
5.3 Benefit-cost relations of interacting with alien bees 73
5.4 Case study 1: Bumble-bee invasion and raspberries 74
5.5 Case study 2: Coffee and the African honey bee in the Americas, revisited 76
Advances in Ecological Research, Volume 63 #2020 Elsevier Ltd
ISSN 0065-2504 All rights reserved.
6. The future of agriculture in a context of bee invasions 79
Increasing honey demand and global coverage of pollinator-dependent crops within
the context of global pollinator declines have accelerated international trade in
managed bees. Bee introductions into agricultural landscapes outside their native
ranges have triggered noteworthy invasions, especially of the African honey bee in
the Americas and the European bumble bee Bombus terrestris in southern South
America, New Zealand, Tasmania, and Japan. Such invasions have displaced native bees
via competition, pathogen transmission, and invaders’capacity to exploit anthropo-
genic landscapes. At high abundance, invasive bees can degrade the mutualistic nature
of many of the flower-pollinator interactions they usurp, either directly by affecting
flower performance or indirectly by reducing the pollination effectiveness of other
flower visitors, with negative consequences for crop pollination and yield. We illus-
trate such effects with empirical examples, focusing particularly on interactions in
the Americas between B. terrestris and raspberry and between the African honey bee
and coffee. Despite high bee abundance and flower visitation in crops, theoretical
and empirical evidence suggests that agricultural landscapes of pollinator-dependent
crops dominated by invasive bees will be less productive than landscapes with more
diverse pollinator assemblages. Safeguarding future crop yield and aiding the transition
to more sustainable agricultural landscapes and practices require we address this
impact of invasive bees. Actions include tighter regulation of the trade in bees to dis-
courage further invasions, reducing invasive bee densities and dominance, and active
enhancement of ecological infrastructure from field to landscape scales to promote
wild bee abundance and diversity for sustained delivery of crop pollination services.
Interaction between human activities and natural ecosystem processes
increasingly determines the characteristics of contemporary landscapes. In
particular, food production by conventional intensive agriculture (Vanbergen
et al., 2020) is a major modifier of global landscapes (IPBES, 2019). This activity
involves land transformation through industrial-scale management of livestock
and crop monocultures with large external inputs and mechanization (Kova
´nszki et al., 2017;Pretty, 2018). The widespread adoption of this model
of food production has simplified agroecosystems by homogenizing land-
scape structure, biotic composition, and ecosystem functioning (Dainese
et al., 2019;IPBES, 2019;Kova
´nszki et al., 2017;Potts et al.,
50 Marcelo A. Aizen et al.
2016;Rusch et al., 2016). As one example, agricultural intensification neg-
atively affects native pollinators (Vanbergen and The Insect Pollinators
Initiative, 2013), diverse animals that provide substantial benefits to humans
through a web of interactions with other organisms, especially by providing
pollination services to crops (Potts et al., 2016).
Paradoxically, global agriculture increasingly depends on pollinators,
owing to greatly intensified cultivation of fruit and seed crops for which
animal pollination enhances yield (Aizen and Harder, 2009). The percentage
of agricultural land planted with pollinator-dependent crops grew world-
wide from 19.4% in the early 1960s to 32.8% by 2016 (Aizen et al., 2019a).
This increase largely reflects a 30% expansion of the global area of agricul-
tural land, mostly dedicated to crops that depend on pollinators (Aizen and
Harder, 2009;Aizen et al., 2008a;Garibaldi et al., 2011a), especially trans-
genic oil-seed crops grown in monocultures (e.g. soybean, canola, and oil
palm) and a diversity of temperate and tropical fruit crops (Aizen et al.,
2019a). The spread of pollinator-dependent crops correspondingly increases
the need for the ecological services provided by wild and domesticated
pollinators (Aizen and Harder, 2009), primarily but not exclusively bees
(Rader et al., 2016), to maximize yield (Potts et al., 2010).
Just as demands for crop pollination are rising, one consequence of this
agricultural expansion and intensification has been the severe depletion of
nesting and floral resources (Baude et al., 2016;Biesmeijer et al., 2006;
Scheper et al., 2014) that support the health and survival of wild pollinator
populations (Goulson et al., 2015;Kennedy et al., 2013;Potts et al., 2010,
2016;Vanbergen et al., 2018). Indeed, landscape simplification and homog-
enization, along with intensive use of pesticides associated with large-scale
agriculture, are among the leading causes of a worldwide pollinator decline
(Brittain et al., 2010;Potts et al., 2010;Vanbergen and The Insect Pollinators
Initiative, 2013;Zattara and Aizen, 2019). Such effects compromise agricul-
tural productivity because visitation by wild bees enhances yield in most
pollinator-dependent crops (Brittain et al., 2013a;Garibaldi et al., 2013;
Potts et al., 2016;Rader et al., 2016). As a consequence, agriculture increas-
ingly relies on pollination by “domesticated” or “managed” bees (IPBES,
2016;Klein et al., 2007;Morse and Calderone, 2003;Pitts-Singer and
Cane, 2011). Historically, the western honey bee (Apis mellifera) was the
managed bee of choice, and it remains so for many crops. However, honey
bees pollinate some crops poorly (e.g. alfalfa, tomato, peppers) and cannot
be used effectively in confined conditions such as greenhouses (Bosch and
Kemp, 2002;Velthius and van Doorn, 2006). Therefore, methods have
51Invasive bees and agriculture
been developed for large-scale rearing and maintenance of other bees, espe-
cially bumble bees (Bombus spp.) and some solitary bees (e.g. Megachile
These agricultural trends along with increasing demand for honey by
a growing human population (Aizen and Harder, 2009) have greatly
boosted worldwide trade in bees (IPBES, 2016), accelerating the transport
of bees within and well beyond their native ranges. Some bees, notably
A. mellifera and Bombus terrestris, subsequently became invasive: dispersing,
surviving at high abundance, and reproducing in a wide range of habitats
far beyond their introduction sites (Morales et al., 2017). Even though
bee trade may bring short-term benefits, these invading species have had
diverse long-lasting negative effects, including impacts on the abundance
and composition of native pollinator assemblages from local to continental
scales (Aizen and Feinsinger, 1994a;Bommarco et al., 2012;Chacoff and
Aizen, 2006;Goulson, 2003;Herrera, 2020;Mallinger et al., 2017;Morales
et al., 2013;Zattara and Aizen, 2019). Given that 88% of angiosperm species
are animal-pollinated (Ollerton et al., 2011) and seed production of over 70%
of crop species benefits from animal pollination (Klein et al., 2007), especially
by wild bees (Garibaldi et al., 2013), bee invasions pose considerable risks for
populations of wild and crop plants (Pottsetal.,2016).
Here, we describe and anticipate the possible impacts of invasive bees on
plant reproduction (seed set and fruit production), with emphasis on yield
in pollinator-dependent crop species. Although these bees can enhance crop
production when either properly managed or in agricultural landscapes
devoid of alternative pollinators, we propose that they can pose a hazard
for both native biotas and agriculture when they become too abundant or
To set the context, we first outline a benefit-cost model of how high
abundance of invasive bees can alter original plant-pollinator interactions,
affecting seed and fruit output (Aizen et al., 2008b;Morales et al., 2017;
Morris et al., 2010). We then review the principal invasions precipitated
by growing worldwide trade in bees, focusing on the relatively recent, geo-
graphically extensive invasions of an African subspecies of the western honey
bee (A. m. scutellata) and the European buff-tailed bumble bee (B. terrestris)of
the Americas (Morales et al., 2017). Although bee trade initiated these
invasions, we also discuss the role of local conditions that facilitated estab-
lishment and range expansion of these bees once introduced, including
the existence of ecological opportunities—perhaps resulting from the
displacement of native bees—and anthropogenic habitat disturbance.
52 Marcelo A. Aizen et al.
Together, the conceptual framework, historical accounts of invasions, and
understanding of local drivers of invasion reveal scenarios of impacts of
introduced bees on crop yield.
We then explore these impacts from several perspectives. First, we
review the importance of A. mellifera and B. terrestris as agricultural pollina-
tors in terms of the number of crops they pollinate. Next, we examine the
abundance and visitation frequency of these species when they become inva-
sive, as the benefit-cost model predicts reduced crop yield when invasive
species become too abundant or visit too excessively. Finally, we assess
evidence for decreases in seed production and crop yield arising from
increasing interaction costs at high visitation frequency. We specifically
use the benefit-cost model to interpret two case studies involving reduced
fruit quality in commercial raspberry fields affected by the invasion of
southern South America by B. terrestris and Roubik’s (2002) proposal that
invasion of the Neotropics by African honey bees improved coffee yield
by increasing pollination stability.
We close by discussing the future of agricultural landscapes within the
context of bee invasions. We argue that the principles revealed by bee inva-
sions apply also to the consequences of increasing simplification of bee
assemblages and their domination by a few species caused by anthropogenic
disturbance, in general, and intensive agriculture, in particular.
2. Plant-pollinator interaction as a benefit-cost relation
Most animal pollinators inadvertently transport pollen within and
among plants while foraging for resources or searching for other opportu-
nities that flowers provide. Plants benefit from this interaction to the extent
that conspecific dispersal improves siring success and fruit and seed produc-
tion. Most flower-visiting animals benefit from available food, especially
sugar in nectar and/or protein in pollen, but also oils, developing seeds,
and “food bodies” (Simpson and Neff, 1981;Willmer, 2011). In some cases,
flower visitors instead seek pheromone precursors, resins for nest con-
struction, heat, shelter or mating sites. Costs associated with this interaction
include the energy and resource investment in floral “rewards” and signals
by plants (e.g. Andersson, 1999;Ashman and Schoen, 1994;Harder and
Barrett, 1992), and the energy and time expended on searching numerous
flowers for limited gain per flower by animals (e.g. Chittka and Spaethe,
2007;Schmid-Hempel et al., 1985) while increasing predation risk
53Invasive bees and agriculture
(e.g. Dukas and Morse, 2003). When gross benefits exceed costs for both
partners, plants and pollinators, the result is mutualism, which is the
prevalent characterization of the plant-pollinator interaction.
Because each partner in a mutualism acts in its own self-interest (i.e.
maximizing individual net benefit), improved performance by one partner
can increase costs for its counterpart (Bronstein, 2001). Indeed, under
certain conditions a flower-animal interaction can instead be antagonistic
(i.e. costs exceed benefits), such as deceitful pollination ( Johnson and
Schiestl, 2016), nectar robbing (Irwin et al., 2010) or pollen theft
(Hargreaves et al., 2009). One such condition is extreme abundance of
one of the partners relative to the other, which can increase interaction costs
via excessive interaction frequency (Morris et al., 2010), as often observed in
plant-pollinator interactions involving invasive bees (Aizen et al., 2014;
Morales et al., 2017). Interaction costs can arise either from direct interac-
tion of the partner species or indirectly as mediated by a third mutualistic
species, for instance a second pollinator (Fig. 1). From a plant’s perspective,
these density-dependent direct and indirect interaction costs are most
relevant to the impact of invasive bees on seed production and, by extension,
to their impact on crop yields in future landscapes.
2.1 Direct costs
Interactions of animals with flowers can impose diverse direct costs on
plant reproduction (Table 1). These costs, such as nectar robbing and flower
damage, can accumulate as individual flowers receive more visits (Aizen
et al., 2014;Morris et al., 2010), as influenced by pollinator abundance
(Garibaldi et al., 2020). Extreme pollinator abundance, such as that observed
for some invasive bees or in fields stocked with managed bees at high
densities (Garibaldi et al., 2020;Sa
´ez et al., 2014), can translate into uncom-
monly frequent pollinator visits per flower, exacerbating costs (Fig. 1A).
Aggravated direct interaction costs to plants as pollinator abundance
increases can cause a shift from mutualism to antagonism (Fig. 1A). The
benefits of interaction, namely ovule fertilization leading to seed production
and seed siring, generally increase with pollinator visitation when visits are
infrequent. As visitation increases, however, seed production reaches an
asymptote either because all ovules are fertilized or because insufficient
resources are available to develop additional seeds. This female limit can
be reached after only a few visits per flower, even though more visits
might benefit pollen export (Bell and Cresswell, 1998;Carlson, 2007).
54 Marcelo A. Aizen et al.
Fig. 1 Expected effects of increased abundance of an invasive bee species (bumble
bee symbol) on benefits of pollinator interaction for plant reproduction when the
invasive species imposes direct (A) or indirect (B–D) interaction costs. The left- and
right-hand scenarios, respectively, illustrate the nature of interactions at low and high
density of the invasive species, which are linked by a blue arrow. For each scenario, the
sign above a black arrow indicates the qualitative interaction effect and arrow width
represents the quantitative effect. Solid and dotted arrows depict direct and indirect
effects, respectively. Panel (A) illustrates a shift from high to low benefit-cost ratios
for the plant associated with aggravated direct interaction costs as invader density
increases. Panels (B–D) show three means by which indirect effects of intensified com-
petition between invading and native bees could weaken mutualistic benefits for the
plant, including: (B) decreased per-visit pollination effectiveness of native pollinators;
(C) replacement of effective native pollinators by less effective alien pollinators;
or (D) decreased diversity of native pollinators.
Beyond this visit intensity, direct plant costs arising from increasing frequen-
cies of animal visitation per flower can increase linearly or even accelerate
(Aizen et al., 2014;Morris et al., 2010). As a result, net benefits, such as fruit
production and production of seeds per fruit, should exhibit a hump-shaped
relation to visit frequency (Fig. 2A) and could even reduce seed set below
that resulting from autonomous self-pollination. Such a relation has been
both predicted for raspberry fields invaded by B. terrestris in Patagonia based
on saturation of pollination benefits accompanied by continually increasing
Table 1 Direct interaction costs and their potential impacts.
Cost Potential impact References
Increased photosynthate expenditure
on nectar production
Ordano and Ornelas
(2005) and Pyke (2016)
Nectar robbing Increased pollination limitation of
seed production due to fewer visits
by legitimate pollinators
Irwin et al. (2010)
Pollen theft Decreased pollen available for export
Hargreaves et al. (2009)
Flower damage Decreased flower attractiveness when
the corolla is damaged, decreased
ovule fertilization when the stigma
or style is broken, or decreased seed
development when the ovary is
´ez et al. (2014) and
Traveset et al. (1998)
Reduced stigmatic pollen load Gross and MacKay
and Young and Young
Failure of pollen-tube populations in
overcrowded styles due to scramble
Harder et al. (2016a,b)
Sexual castration Jennersten (1988) and
Shykoff and Bucheli
Altered nectar composition deterring
Herrera et al. (2013),
Vannette et al. (2013)
and Vannette and
56 Marcelo A. Aizen et al.
flower damage costs (Sa
´ez et al., 2018) and demonstrated in a meta-analysis
of studies on crops subject to high visitation (≫10 visits flower
managed honey bees (Rollin and Garibaldi, 2019; see Section 5).
Under most natural circumstances, plants influence the number of
pollinator visits received by individual flowers through the genetic determi-
nation of floral lifespan and attractiveness (Ashman, 2004;Klinkhamer et al.,
1993;Pyke, 2016) and plastic responses to actual visit frequency (Harder and
Johnson, 2005;van Doorn, 1997). Extremely frequent visitation that dam-
ages flowers or leads to unusually large stigmatic pollen loads that promote
massive pollen-tube abortion (Table 1) is rare in nature. As a consequence, a
hump-shaped relation of plant reproductive output to visit frequency is
likely uncommon and difficult to distinguish from a saturating relation under
most natural situations, for which interaction costs do not typically exceed
interaction benefits (Harder et al., 2016a,b). Nevertheless, adaptive regula-
tion of pollinator visitation by different floral traits may fail when a pollinator
becomes super-abundant, generating a hump-shaped relation. This conse-
quence seems to arise when invasive bees reach locally high densities
(Aizen et al., 2014).
2.2 Indirect costs
Massive invasions by alien pollinators can affect pollination by native polli-
nators indirectly through negative effects on the pollination efficiency/
effectiveness, abundance, and/or diversity of native species (Fig. 1B–D).
Fig. 2 Predicted general relations of seed output to visitation frequency by an invasive
bee species if increased invader density primarily exacerbates (A) direct interaction
costs or (B) indirect costs.
57Invasive bees and agriculture
Such effects could arise from competition between alien and native bees for
floral resources or nest sites or from introduction of pathogens by invading
bees (Goulson, 2003). In either case, seed production will decline with
increasing visit frequency by the invader if reduced pollination by native pol-
linators is not offset by increased pollination by the invasive pollinator, raising
the indirect cost of a plant’s overall interaction with pollinators (Fig. 2B).
The quantity and quality of pollen deposited on a flower’s stigma per visit
varies widely among pollinators (Herrera, 1987, 1989). From the perspec-
tive of a plant’s female function, deposition of compatible pollen per visit is a
relevant measure of an animal’s per visit pollination effectiveness, whereas
the product of mean deposition of compatible pollen during individual visits
and the frequency of visits is a relevant measure of overall pollination effec-
tiveness (Willmer, 2011). Pollen deposition per visit by an effective pollinator
and subsequent seed production can decline with increasing visit frequency
of a less effective floral visitor (Fig. 1B). For example, bees that actively
collect pollen to provision their larvae can greatly decrease pollen availability,
so pollen transport to stigmas by another, otherwise effective pollinator can be
reduced or even eliminated in the presence of pollen thieves (Chalcoff et al.,
2012;Hargreaves et al., 2009, 2010). Likewise, by reducing the availability
of reward a nectar robber can shorten visits by legitimate pollinators, reduc-
ing both pollen removal and deposition and possibly inducing otherwise
legitimate pollinators to engage in secondary robbing (Irwin et al., 2010).
Given a constant total number of visits, displacement of an effective
native pollinator by a less effective alien pollinator could reduce the total
quantity or quality of pollen deposited on stigmas of a focal plant species
(Dohzono and Yokoyama, 2010;Fig. 1C). Such displacement could arise
from foraging responses, if native pollinators react to resource competition
by shifting their attention to other plant species. For instance, increased vis-
itation by an invasive nectar robber can render flowers unrewarding for a
legitimate pollinator and decrease its visit frequency (Irwin and Brody,
1998). Invasive social bees, like the African honey bee, which tend to pro-
vide low-quality pollination by visiting sequentially several flowers per
plant, can also reduce visitation by high-quality pollinators that move more
pollen among plants (e.g. Aizen and Feinsinger, 1994a,b). Displacement can
become permanent when competition or pathogen transmission depletes or
extirpates populations of native pollinators (e.g. Arbetman et al., 2012;
Morales et al., 2013).
Displacement of native pollinators by an invader will also generally
reduce the species diversity of a plant’s pollinators. Plants often benefit in
58 Marcelo A. Aizen et al.
various ways from interacting with a diverse pollinator assemblage. First,
different pollinators can forage in different habitats or microhabitats where
a plant species occurs and blooms, or be active at different periods of the
day or flowering season, thus providing more stable and effective pollination
service (Albrecht et al., 2012;Brittain et al., 2013a). Second, the probability
of visitation by an effective pollinator species increases with pollinator diver-
sity (Albrecht et al., 2012). This sampling effect is particularly relevant
agriculturally because crops typically flower for short periods and flowering
phenologies vary spatially and temporally (Bartomeus et al., 2013;Winfree
et al., 2018). Last, antagonistic interactions between pollinator species
can increase pollination quality. Interference or exploitation competition
for floral resources can induce pollinators to fly farther between consecu-
tive visits, promoting outcrossing (Brittain et al., 2013b;Greenleaf and
Kremen, 2006). Combining all these influences, reduced diversity of a
pollinator assemblage in the presence of a highly abundant, dominant,
managed, or invasive pollinator can decrease both pollination quantity
and quality (Fig. 1D).
3. Invasive bees
Range expansion occurs naturally during the histories of most species
(Chuang and Peterson, 2016;Tomiolo and Ward, 2018). However, human-
aided expansion often has a different character than this natural process,
in part because it commonly involves establishing populations far beyond
a species’ native range (Blackburn et al., 2014;Hastings et al., 2005). The
establishment success of these populations can be augmented by repeated
human “seeding” (Simberloff, 2009) that rescues them from early extinction
(Brown and Kodric-Brown, 1977). Many bee species have established out-
side their native ranges following accidental or intentional introductions
(Goulson, 2003). For instance, cavity-nesting Anthidium manicatum was acci-
dentally transported from Europe to North America along with its nesting
material, whereas other cavity-nesting Megachile and Osmia have been trans-
ported intentionally in the same direction to improve pollination services
(Gibbs and Sheffield, 2009). Several species introduced both accidentally
and intentionally have established self-sustaining, expanding populations
in the new regions to which they were transported (Aizen et al., 2019b;
Geslin et al., 2017b;Goulson, 2003;Morales et al., 2017)—i.e. they became
invasive (sensu Pys
ˇek and Richardson, 2006).
59Invasive bees and agriculture
3.1 The western honey bee
The western honey bee, A. mellifera, is one of eleven Eurasian Apis species
that are characterized by their perennial eusocial colonies (Engel, 1999). Its
native range extends from northern Europe to southern Africa and from
western Europe to the Ural Mountains and the Arabian Peninsula. More
than 25 subspecies are recognized throughout its distribution (Schneider
et al., 2004). For more than 4000 years, long before the pollination role of
honey bees was appreciated, humans have managed this species as a source
of honey, a derivative of floral nectar that colonies store to survive inclement
periods (Crane and Graham, 1985). Because honey bee colonies can be
maintained and moved easily in artificial containers, they have been trans-
ported from their native range to every continent except Antarctica (Moritz
et al., 2005). During the 20th century, honey bees also became the species
of choice for intentionally enhancing crop pollination (IPBES, 2016).
In landscapes permitting year-round survival, honey bee swarms from
managed colonies often establish feral colonies and become self-sustaining
components of the local bee (and pollinator) fauna (Hung et al., 2018),
often with detrimental consequences for native species (Mallinger
et al., 2017).
Apis mellifera ligustica is the most widespread managed honey-bee
subspecies, but it performs poorly under tropical conditions (Ruttner,
1988), which motivated the search for other subspecies to enhance honey
production in tropical countries. In 1956, queens of the African subspecies
A. m. scutellata were imported to Sao Paulo State, Brazil, to hybridize
with A. m. ligustica (Moritz et al., 2005). The subsequent accidental release
of 26 swarms of African honey bees in 1957 started the most exten-
sive bee invasion to date. By the mid-1970s the American range of this
bee extended throughout most of the central and eastern Amazonian
Basin and south to central Argentina. African honey bees were first observed
in Central America during 1982, reached Mexico in 1985 and arrived in the
southern USA during the 1990s (Moritz et al., 2005). The current range of
the African honey bee in the Western Hemisphere covers >15,000,000 km
including most of South and Central America and southern states of
the USA. Throughout its range, the African honey bee has remained
largely distinct genetically from other European honey bee subspecies
(Schneider et al., 2004;Smith, 1991). Given this limited hybridization
and introgression, we refer to this invasive bee as African, rather
than Africanized as in some other accounts of this invasion (Morales
et al., 2017).
60 Marcelo A. Aizen et al.
3.2 Bumble bees
Bumble bees, Bombus spp., comprise about 260 eusocial species of large bees
(Williams, 1998). They are mostly cold-adapted and, unlike honey bees,
their colonies are annual, except for those of a few tropical species, like
B. pauloensis (Sakagami, 1976). Eurasia has the most diverse Bombus fauna,
followed by North America and South America (Williams, 1998). No native
bumble bees occur in Oceania or Africa south of the Mediterranean
coast. Bumble bees are important pollinators of many crops, especially those
requiring active vibration of the anthers to release pollen (buzz-pollination),
under field and greenhouse conditions (Velthius and van Doorn, 2006).
Methods have been developed for rearing and transporting several bumble
bee species. Two species in particular, the European B. terrestris and the
North American B. impatiens, are reared at industrial scales for managed
pollination (Reade et al., 2015). Although one-third of Bombus species
are declining, some species, particularly those that are managed, are thriving
(Arbetman et al., 2017).
Bumble bee species have been translocated globally for agricultural pol-
lination. Four species, B. hortorum,B. terrestris,B. subterraneus and B. ruderatus,
were first introduced and established in New Zealand from the UK more
than a century ago (Howlett and Donovan, 2010;Macfarlane and Gurr,
1995). About 300 queens of long-tongued B. ruderatus were shipped from
New Zealand to Chile during the early 1980s (Arretz and Macfarlane,
1986). Short-tongued B. terrestris hasbeenintroducedfromEuropeto
Israel, northern Africa, Asia, Central America and Chile, and secondarily
from New Zealand to Japan and Tasmania (Goulson, 2003;Montalva
et al., 2011). Some of these introductions resulted in extensive invasions
(Aizen et al., 2019b). Bombus ruderatus became invasive in southern South
America and its range now extends along both sides of the Andes to more
than 400km south of the original introduction sites in south-central Chile
(Morales et al., 2013). This invasion was eclipsed by that of B. terrestris,which
has invaded every region into which it has been introduced, including
New Zealand, Tasmania, Japan, and South America. This remarkable inva-
sive potential is most evident in South America. Since 1997, B. terrestris has
expanded its South American range more than 2000km southward from the
original introduction sites in central Chile, to the southernmost islands of
the continent (Cape Horn, south of Tierra del Fuego), and from the
Pacific to the Atlantic coasts of Patagonia (Aizen et al., 2019b). Abiotic niche
models predict further range expansion by B. terrestris northward to southern
´, Bolivia and Brazil, and across the Argentine Pampas (Acosta et al., 2016).
61Invasive bees and agriculture
3.3 Other bees
Several species of solitary bees, mostly Megachilidae, have been intentionally
introduced beyond their native ranges for pollination, largely owing to their
capacity to trigger the explosive release of the concealed stamens and pistil
of alfalfa flowers, which is required for pollination. The most extensive
introduction involves the European Megachile rotundata, introduced in large
numbers to South and North America for alfalfa pollination and now reared
and traded extensively in the USA and Canada (Pitts-Singer and Cane,
2011). Several Osmia species have also been introduced to the Americas
from Europe and Asia to pollinate temperate fruit crops. Osmia cornuta
was introduced to California from Spain for almond pollination, whereas
O. cornifrons was introduced to the east coast of the USA from Japan for apple
pollination (Goulson, 2003, and references therein). Osmia species have also
been introduced to non-native regions within countries. For example,
O. ribifloris biedermannii was introduced from the west to the east coast of
the USA for blueberry pollination (Stubbs et al., 1994). A halictid example
involves Nomia melanderi, which was introduced in New Zealand from
North America during the early 1970s for alfalfa pollination (Howlett and
Donovan, 2010). Although some introduced solitary bees have established
self-sustaining feral populations and even expanded their geographical
ranges (Goulson, 2003), they do not appear to have dominated native
pollinator communities or have had noteworthy effects on the native bee
fauna in the invaded ranges (see Table 2.4.3 in IPBES, 2016). For instance,
the long-term decline of native O. lignaria in eastern North America appears
to have been unrelated to the introduction and range expansion of its alien
congeneric O. cornifrons in the late 1970s (Centrella, 2019). Although these
examples suggest that introduced solitary bees have limited effect on native
bee faunas, this remains to be assessed by targeted impact assessment.
4. Drivers of bee invasion success
Whether humans are involved or not, species invasions, including bee
invasions, encompass three principal processes: arrival of a species to a novel
geographical area, initial establishment, and subsequent spread (Davis,
2009). Species sometimes disperse long distances without human aid, such
as crossing oceans by drifting on air or water currents or by arriving on
dispersal agents such as migratory birds (Nathan et al., 2008;Nogales
et al., 2012). Once a species reaches a new location, among other factors,
62 Marcelo A. Aizen et al.
its fate depends on whether the local environment permits establishment,
as determined by abiotic suitability, the relative abundance of facilitating
species and the relative rarity of competitors and enemies (Davis, 2009).
If these conditions exist, newly arrived species may also displace species
occupying similar niches (Catford et al., 2018). Once established, the
species’ subsequent spread depends on dispersal ability and increased
propagule pressure provided by local production and external input of
dispersing individuals (Davis, 2009;Simberloff, 2009). Humans contribute
to all aspects and stages of species invasion by inadvertently or intentionally
transporting thousands of species, often repeatedly, to new ranges ( Jeschke
and Strayer, 2005;Meyerson and Mooney, 2007), and by creating condi-
tions favourable to their establishment and population growth through
habitat degradation and suppression of competitors or enemies (Davis,
2009;Lodge, 1993;Shea and Chesson, 2002). Therefore, trade, ecological
opportunities and displacement of native counterparts, and anthropogenic
disturbance have probably all facilitated bee invasions.
4.1 Bee introductions and trade
Bee introduction, a necessary but not sufficient condition for the occurrence
of a bee invasion, has become increasingly frequent and extensive in recent
years. In addition to the multiple unintended or intended introductions
of small numbers of individual bees (e.g. bumble bees in New Zealand,
Howlett and Donovan, 2010; African honey bees in Brazil, Moritz et al.,
2005), millions of bees have been produced and traded within and among
continents since the 1980s (Aizen et al., 2019b;Armitage, 2018;Owen,
2017;Velthius and van Doorn, 2006). This burgeoning international bee
trade is both economically profitable and a major threat to biodiversity
(Sutherland et al., 2017). Most recent trade in bees involves A. mellifera
reared in Australia (Armitage, 2018) and B. terrestris reared in Europe and
Israel (Aizen et al., 2019b). Also reared commercially are other bumble
bee species, such as B. impatiens in eastern North America, B. ignitus in
Japan, and more recently B. pauloensis in South America, as well as several
species of alkali, mason and leaf-cutter bees. Apart from Megachile rotundata
and to a lesser extent a few Osmia species (Goulson, 2003), no other bee
species have been involved in transcontinental trade at a scale compar-
able to that of A. mellifera and B. terrestris (Aizen et al., 2019b;Geslin
et al., 2017b;Goulson, 2003, 2010;Morales et al., 2017;Stout and
63Invasive bees and agriculture
In general, species introduction can involve one of three geographical
and biological scenarios. One involves introduction of a subspecies where
another subspecies of the same species is already resident, whether naturally
or owing to a previous introduction. Examples include translocations of
European subspecies of the western honey bee and of bumble bee species
within Europe or North America (Bartomeus et al., 2020;Elie, 2015;
Goulson, 2010) and the introduction of the African honey bee, A. m.
scutellata, into the Americas, where other European subspecies had already
been introduced (Schneider et al., 2004;Smith, 1991). The second scenario
involves introduction where native congeners were already present, such as
the introductions of the western honey bee into eastern and southern Asia
and of B. terrestris in eastern Asia and South America (Goulson, 2003, 2010).
The final scenario involves introduction of a bee species into a region not
previously occupied by any congener, such as the original introductions
of A. mellifera in the Americas and Oceania and of different of Bombus species
in New Zealand (Howlett and Donovan, 2010;Macfarlane and Gurr, 1995;
Moritz et al., 2005). These different types of introduction have initiated
minor and major bee invasions (see Section 3).
Intentional species introductions, such as those motivated by commercial
benefit, are much more likely to cause and maintain ecologically significant
invasions if these species are continually introduced in large numbers. The
rate and extent of invasion spread depends on propagule pressure—i.e. the
number of individuals released in an unoccupied area (Lockwood et al.,
2005;Simberloff, 2009). Accordingly, continuous supply of bees due to
well-established, unregulated bee imports can subsidize and accelerate
ongoing invasions. Such augmentation appears to enhance the invasion of
southern South America by B. terrestris. Since 1997, more than 1.2 million
colonies and queens have been continually imported into Chile, the only
country in South America allowing this trade, from bumble-bee factories
in Europe and Israel for greenhouse tomato pollination and open-field pol-
lination of blueberry and other crops (Fig. 3:Aizen et al., 2019b;Montalva
et al., 2011). Although B. terrestris is now well established in the wild in Chile
and Argentina, and probably in Bolivia and Peru
´, each new importation
could result in further introgression from a genetically novel propagule stock
and introduction of new pathogens spilling over to the native bees, which
in turn could promote the ongoing invasion (Smith-Ramı
´rez et al., 2018).
For similar reasons, trade of a bee species within its native region can have
invasion-like consequences by artificially increasing the numerical and
ecological dominance of the supplemented species. This effect likely
64 Marcelo A. Aizen et al.
contributes to the dominance of A. mellifera and perhaps of B. terrestris
among European bee assemblages (Bommarco et al., 2012;Goulson,
2010;Herrera, 2020). The consequences of growing bee trade for propagule
pressure have been recognized as a key factor promoting bee invasions by the
Intergovernmental Science-Policy Platform on Biodiversity and Ecosystem
Service (IPBES, 2016).
4.2 Ecological opportunities and displacement of native bees
Once an alien species has been introduced, ecological characteristics of
the new geographical area and its resident species determine whether the
newcomer establishes and spreads (Davis, 2009). In particular, invasion is
promoted by the existence of ecological opportunities, including the
absence of ecologically similar competitors (lack of biotic resistance; Shea
and Chesson, 2002), and of predators or pathogens (enemy-free space;
Fig. 3 Numbers of colonies and queens of Bombus terrestris imported into Chile from
1997, when it was first introduced, through June 2016. Data from the Servicio Agrícola
Ganadero of Chile (SAG, 2016. Data Provided Upon Request in June 2016 According to
the Law of “Transparency and Access to Public Information”[Reference AR006T0000668].
Servicio Agrícola Ganadero of Chile, Santiago, Chile). Adapted from Aizen, M.A., Smith-
Ramírez, C., Morales, C.L., Vieli, L., Sáez, A., Barahona-Segovia, R.M., Arbetman, M.P.,
Montalva, J., Garibaldi, L.A., Inouye, D.W., Harder, L.D., 2019b. Coordinated species importa-
tion policies are needed to reduce serious invasions globally: the case of alien bumblebees
in South America. J. Appl. Ecol. 56, 100–106.
65Invasive bees and agriculture
Shea and Chesson, 2002), and/or the availability of mutualists (Davis, 2009).
For example, successful, intentional introduction of four bumble bee
species to New Zealand during the late 19th century benefited from
both the natural absence of a potentially competing long-tongued bee fauna
(Anderson, 2003) and human cultivation of mutualists in the form of non-
native leguminous forage crops adapted to pollination by long-tongued
bees (Goulson, 2003;Macfarlane and Gurr, 1995). The latter role of non-
native mutualisms reflects a broader phenomenon of “invasion complexes”,
whereby alien pollinators disproportionally visit alien plants for nectar
and pollen, pollinating the plants visited and promoting their reproduction
(Morales and Aizen, 2002, 2006). Thus, plant and pollinator invasions can
reinforce each other.
Even if related species occur in the new habitat, an invasion can
succeed if the introduced species is competitively superior to residents,
perhaps benefiting from the mutualistic interactions previously shaped by
functionally similar residents over evolutionary time (Shea and Chesson,
2002). Indeed, most studies of the effects of invasive and managed bees
on native bees have detected negative effects (Mallinger et al., 2017).
Relevant traits of highly invasive bees like A. m. scutellata and B. terrestris
that make them superior competitors and often dominant species in their
native home ranges include sociality, diet generalization, efficient foraging,
and aggressiveness (Bommarco et al., 2012;Geslin et al., 2017b;Schneider
et al., 2004;Smith, 1991;Velthius and van Doorn, 2006). For example, the
rapid and extensive spread of African honey bees in the Americas compared
to other established A. mellifera subspecies resulted in part from faster
colony growth aided by greater attention to pollen foraging, and in part
from usurpation of established nests of European subspecies (Schneider
et al., 2004). Competition for food and nest sites has also been implicated
in the decline of Apis cerana in China after the introduction of European
A. m. ligustica in 1896 (Yang, 2005). Bombus terrestris likely benefits from a
different competitive advantage as it continues displacing the only native
bumble bee in southern South America, B. dahlbomii.AlthoughB. terrestris
has a shorter proboscis than B. dahlbomii, it commonly robs long-tubed
flowers of species previously pollinated by B. dahlbomii, thereby depleting
the native species’ nectar sources (N.M. Rosenberger, M.A. Aizen and
L.D. Harder, unpublished data).
Invaders can also displace residents via apparent competition, particularly
if they are vectors of pathogens that they tolerate but that are detrimental to
resident species (Holt and Bonsall, 2017). Pathogen transmission from
66 Marcelo A. Aizen et al.
commercial colonies of honey bees and bumble bees may be a leading con-
tributor to the global pollinator decline (F€urst et al., 2014;IPBES, 2016;
Vanbergen and The Insect Pollinators Initiative, 2013;Vanbergen et al.,
2018). For instance, introduction of colonies of invasive B. terrestris in
South America also introduced the highly pathogenic protozoan Apicystis
bombi, which infected B. dahlbomii and the previously introduced alien
B. ruderatus (Arbetman et al., 2012). Furthermore, introduced pollinators
carrying multiple pathogens can increase their infectious potential because
of complex synergistic interactions. For instance, the host-shift by ectopar-
asitic Varroa mites to the western honey bee as a consequence of the global
trade led to increased prevalence and virulence of strains of the deformed
wing virus (DWV) through cycles of transmission during Varroa feeding
and recombination in this honey bee host (Vanbergen et al., 2018).
Furthermore, spillover of pathogenic DWV strains has occurred between
managed honey bees and wild Bombus species (F€urst et al., 2014;Wilfert
et al., 2016). Consequently, introduced and managed bees can facilitate
the spread of new diseases to wild populations, eliciting novel epidemics
that may affect population and community structure, with implications
4.3 Anthropogenic disturbance
In contrast to many plant invasions, the spread of introduced bees does
not appear to depend on natural or anthropogenic habitat disturbance.
For example, the African honey bee and B. terrestris have spread impressively
in the Americas by colonizing diverse natural habitats. Nevertheless, these
and other invasive bees (e.g. B. ruderatus) thrive in human-disturbed
habitats (Aizen and Feinsinger, 2003;Morales et al., 2013). For instance,
Aizen and Feinsinger (1994a) found that the abundance of African honey bees
increased with fragmentation of the Chaco forests of NW Argentina, whereas
the abundance and diversity of native bees decreased. In addition, compared
to most other bees A. mellifera is a long-range forager (up to 10km from
the hive; Beekman and Ratnieks, 2000), which may also contribute to its
ubiquity in anthropogenic landscapes. This trait could explain, at least in part,
the results of a meta-analytical study reporting that the abundance of feral
A. mellifera declines more slowly with increasing distance to natural or semi-
natural field margins than that of most native bees (Garibaldi et al., 2011b).
Bombus terrestris can attain high densities in agroecosystems, such as
raspberry fields in southern Argentina (Sa
´ez et al., 2014) and blueberry
67Invasive bees and agriculture
fields in southern Chile (Aizen et al., 2019b). Also, conversion of meadows
and seminatural pastures into arable lands and agricultural intensification
in Sweden have been implicated in the demise of several specialized,
long-tongued bumble bee species and increasing dominance of a generalist,
short-tonged bumble bee such as B. terrestris (Bommarco et al., 2012).
Overall, super-abundant and highly generalist, invasive bee species like
A. mellifera and B. terrestris seem more resistant than native bees to the effects
of habitat homogenization and agrochemicals, which may enable them to
prosper in different disturbed habitats—including agricultural lands—where
native bees do poorly (Aizen and Feinsinger, 2003;Aizen et al., 2014;Arena
and Sgolastra, 2014;Morales and Aizen, 2002).
The greater resistance of invasive bees to different types of anthropogenic
disturbances suggests that they will dominate impoverished bee communi-
ties in homogenous landscapes such as those created by conventional inten-
sive agriculture (Vanbergen and The Insect Pollinators Initiative, 2013).
Indeed, in those landscapes local bee assemblages seem to converge on a
similar set of dominant bee species, independent of the specific anthropo-
genic disturbance, which can reduce αand βdiversity of bees (Aizen and
Feinsinger, 1994a, 2003;Chacoff and Aizen, 2006;Quintero et al., 2010).
On the other hand, high bee diversity can be maintained in heterogeneous
agricultural landscapes that include remnants of natural and seminatural
habitat (Winfree et al., 2007, 2009). Extirpation of many pollinator species
by anthropogenic disturbance, particularly solitary bees, releases floral
resources, including mass-flowering crops, for exploitation by the more
generalist survivors (Steffan-Dewenter et al., 2002;Westphal et al., 2003).
Therefore, large-scale agricultural homogenization of landscapes, while not
a prerequisite for a bee invasion, indirectly favours disturbance-resistant social
bees such as A. mellifera and B. terrestris.
5. Consequences of bee invasions for crop pollination
According to the benefit-cost perspective on plant-pollinator mutu-
alisms elaborated in Section 2, the costs of this interspecific interaction
to plants likely intensifies as bee density, and hence visitation per flower,
increases. Thus, even though invasive bees, and many other bees, can
enhance crop pollination at low or moderate densities, detrimental effects
on agricultural yield are expected if three conditions are met: (1) invasive
bees visit the flowers of many different crops; (2) individual flowers receive
many visits when alien bees become invasive and attain high densities;
68 Marcelo A. Aizen et al.
and (3) very frequent visits reduce a flower’s seed production because of
increased direct or indirect mutualism costs, reducing overall crop yield.
We discuss each of these conditions and review existing evidence from both
crop and wild plants. To illustrate the consequences of invasive bees for crop
yield, we present the case of commercial raspberry and B. terrestris in South
´ez et al., 2014, 2017, 2018) and revisit the proposal that coffee
yield in tropical America increased following invasion by the African honey
bee (Roubik, 2002).
5.1 Interaction of alien bees with crop flowers
Despite the recognized risks of species invasion for pollinators and pollina-
tion (Vanbergen et al., 2018), intentional bee introduction continues to be
justified by the ability of managed bees to produce honey (Apis) and/or their
apparent utility as crop pollinators (Dicks et al., 2016;IPBES, 2016). The
two most-traded bees, A. mellifera and B. terrestris, are extreme generalists,
enhancing their value as managed pollinators of most common pollinator-
dependent crops (see below). Even pollinators considered to be relative
specialists may be gourmands, not gourmets. The alfalfa leaf-cutter bee
(M. rotundata), for example, is increasingly employed to pollinate crops other
than alfalfa and its relatives, such as canola (Westcott and Nelson, 2001).
Apis mellifera is probably the most generalized pollinator of all, visiting the
flowers of thousands of plant species, participating in and, in many cases,
dominating plant-pollinator networks in both its native and introduced
ranges (Hung et al., 2018). Apis mellifera pollinates the flowers of about
80% of pollinator-dependent crops (Klein et al., 2007). Whether managed
or feral, this bee species can dominate the pollinator assemblage of many
tropical and temperate crops (Fig. 4), in some cases accounting for >75%
of all visits to flowers of disparate crops, including apple, coffee, grapefruit,
macadamia, sunflower, and soybean among many others (Badano and
Vergara, 2011;Blanche et al., 2006;Blettler et al., 2018;Chacoff and
Aizen, 2006;Geslin et al., 2017a;Sa
´ez et al., 2012). Apis mellifera’s gener-
alized feeding and extensive foraging range heighten its utility as a crop
pollinator, especially under intensive agricultural management such as
monocultural production of soybean and canola (e.g. Chiari et al., 2005;
de Souza Rosa et al., 2011). For example, visit frequency by African honey
bees to flowers in extensive grapefruit plantations in NW Argentina is only
10% less in the middle of plantations (500m from the edge) than at the
edge, whereas visitation by native bees in the same plantations declined
69Invasive bees and agriculture
85% (Chacoff and Aizen, 2006). Across agricultural landscapes, honey
bees can use the flowers of both crops and of the surrounding vegetation
(Russo et al., 2013).
Bumble bees also effectively pollinate many temperate plant species and
crops. For example, B. terrestris visits flowers of more than half of all temper-
ate animal-pollinated crops considered in a recent meta-analysis (Garibaldi
et al., 2013), including red clover, strawberry, onion, spring rape, turnip
rape, cherry, and fava bean. Another meta-analysis identified this species
as the most important wild pollinator for agriculture in Europe in terms
of the number of crops it pollinates, visitation frequency, and economic
contribution to crop production (Kleijn et al., 2015). Bumble bees are
Fig. 4 Proportion of visits by domesticated and/or feral Apis mellifera to different crops,
ranked in descending order. Data from Table S2 of Garibaldi, L.A., Steffan-Dewenter, I.,
Winfree, R., Aizen, M.A., Bommarco, R., Cunningham, S.A., Kremen, C., Carvalheiro, L.G.,
Harder, L.D., Afik, O., Bartomeus, I., Benjamin, F., Boreux, V., Cariveau, D., Chacoff, N.P.,
offer, J.H., Freitas, B.M., Ghazoul, J., Greenleaf, S., Hipólito, J., Holzschuh, A.,
Howlett, B., Isaacs, R., Javorek, S.K., Kennedy, C.M., Krewenka, K.M., Krishnan, S.,
Mandelik, Y., Mayfield, M.M., Motzke, I., Munyuli, T., Nault, B.A., Otieno, M., Petersen, J.,
Pisanty, G., Potts, S.G., Rader, R., Ricketts, T.H., Rundl€
of, M., Seymour, C.L., Sch€
orgyi, H., Taki, H., Tscharntke, T., Vergara, C.H., Viana, B.F., Wanger, T.C.,
Westphal, C., Williams, N., Klein, A.M., 2013. Wild pollinators enhance fruit set of crops regard-
less of honey bee abundance. Science 339, 1608–1611: see that reference for crop details.
70 Marcelo A. Aizen et al.
particularly useful pollinators of crops with poricidal anthers, such as
tomato and eggplant, because unlike honey bees they can actively produce
the vibrations needed to remove pollen from anthers (buzz-pollination;
Buchmann, 1983). Although most B. terrestris colonies imported to
Chile have been used for tomato pollination, particularly in greenhouses,
they have also been used to pollinate blueberry, avocado, strawberry, canola,
pepper, melon, and broccoli (Estay, 2007). In general, introduced bumble
bees, including B. terrestris, use both alien and native plant species extensively
(Howlett and Donovan, 2010;Montalva et al., 2011).
5.2 Visit frequencies of alien bees
Even though invasive bees may be effective crop pollinators at low densities,
they can be detrimental to agriculture and to plant sexual reproduction
in general when they become extremely abundant (see Section 2). These
effects can arise from increasing direct costs associated with high frequencies
of flower visitation (Fig. 1A, Table 1) or indirect costs arising from extreme
dominance of pollinator assemblages (Fig. 1B–D). The limited available evi-
dence supports the premise that invasive bees commonly become extremely
abundant and dominant (Morales et al., 2017).
The invasive African honey bee is probably the most abundant
bee throughout the Neotropics. For instance, in dry Chaco forest of NW
Argentina the African honey bee accounts for >90% of visits to Prosopis nigra
inflorescences and about 50% of visits to Parkinsonia praecox flowers (Aizen
and Feinsinger, 1994a). A reanalysis of data from Aizen and Feinsinger
(1994a) suggests that African honey bees were respectively two and three
times more abundant visitors to the flowers of these two native tree species
than all other species combined (Fig. 5). Similarly, feral African honey bees
represent >90% of visitors to cultivated citrus flowers in NW Argentina
(Chacoff and Aizen, 2006). These data reveal pollination environments
dominated by invasive African honey bees.
Invasive bumble bees can also attain extreme densities, as illustrated
by examples from successive invasions of the temperate forests of NW
Patagonia by B. ruderatus and B. terrestris (Morales et al., 2017). Bee visits to
Alstroemeria aurea growing in montane Nothofagus forests in the Challhuaco
Valley, Argentina, between 1994 and 2013 provide graphic evidence.
Immediately before the first invasion by alien Bombus,A. aurea flowers
received an average of 0.075 visits h
by the native B. dahlbomii.
Bombus ruderatus first appeared in this forest during 1996, and by 2008
71Invasive bees and agriculture
B. dahlbomii had been extirpated. During 2006, visitation to A. aurea flowers
by B. ruderatus peaked at 0.12 visits h
. The subsequent invasion by
B. terrestris reached the Challhuaco Valley in 2007, and since 2009
A. aurea flowers have received an average of 0.15 visits h
Fig. 5 Frequencies of visits to (A and C) the brush-like inflorescences of Prosopis nigra
and (B and D) individual flowers of Parkinsonia praecox by feral African honey bees (Apis
mellifera scutellata) and native insects in a fragmented dry Chaco forest of NW
Argentina. Panels (A and B) depict visit frequencies for trees isolated in small fragments
(<1ha, red circles), in large fragments (>1 ha, orange circles), and in continuous forests
(yellow circles). Lines represent estimated quantile regressions of variation in median
visit frequency by African honey bees as a function of visit frequency by native insects
(dashed lines) and of visit frequency by native insects as a function of visit frequency by
African honey bees (solid lines). The coloured squares indicate the visitation of each
insect group on its own, as estimated by the regression intercepts for visitation by
African honey bees (blue squares) and by native insects (green squares). Panels
(C and D) compare these intercepts (95% confidence interval) as estimated by the qua-
ntile regressions. Data from Aizen, M.A., Feinsinger, P., 1994a. Habitat fragmentation, native
insect pollinators, and feral honey bees in Argentine ‘Chaco Serrano’. Ecol. Appl. 4, 378–392.
72 Marcelo A. Aizen et al.
the visit frequency by its original native pollinator (Morales et al., 2013;
C.M. Morales, unpublished). More general evidence comes from a survey
conducted during 2011 throughout the Patagonian Andes, when the
native B. dahlbomii still occupied southern landscapes not yet reached by
B. terrestris or B. ruderatus and comparisons between habitats with and
without B. terrestris alone were still possible. The abundance of B. terrestris
in invaded areas was 10 times greater than that of B. dahlbomii in as-yet-
uninvaded areas (Morales et al., 2013). Similarly, raspberry flowers in com-
mercial fields in NW Patagonia receive one to two orders of magnitude more
visits than those in fields in the native range of B. terrestris in Europe (Sa
et al., 2014, 2018). Clearly, bumble bees can become unusually abundant
and visit flowers extremely frequently once they have invaded new regions.
5.3 Benefit-cost relations of interacting with alien bees
Mounting evidence demonstrates a reduction in seed output arising from
direct costs of interacting with highly abundant, dominant pollinators
(Fig. 1A). An extreme example involves fruit set by South American
Capparis atamisquea, which fails almost completely in intensively visited
plants (Morris et al., 2010). In NW Argentina, high visitation by invasive
African honey bees along with the transfer of self-pollen (see below) may
explain decreased fruit set by C. atamisquea in fragmented forests (Aizen
and Feinsinger, 1994b). More broadly, a recent meta-analysis of the effects
of the western honey bee on crop productivity based on 16 crops found
maximum fruit or seed set by flowers that received 8–10 honey-bee visits
during their lives, but reduced seed set by flowers subject to more visits
(Rollin and Garibaldi, 2019). Such peaked relations are consistent with
saturating gross benefits, but linearly increasing costs for female reproductive
success by plants as pollinator visitation increases (Fig. 2A).
Evidence also indicates reduced seed output arising from indirect
costs via decreasing abundance and diversity of other pollinators. In partic-
ular, visit frequency by native insects can vary inversely with that of honey
bees when the latter dominates a floral resource (Aizen and Feinsinger,
1994a;Mallinger et al., 2017;Rollin and Garibaldi, 2019), thereby affecting
plant reproduction (e.g. Aizen and Feinsinger, 1994b;Badano and Vergara,
2011). Pollen-collecting honey bees can decrease per-visit and total polli-
nation by nectar-seeking pollinators (Chalcoff et al., 2012;Hargreaves
et al., 2010), illustrating the indirect costs depicted in Fig. 1B. Also, the
displacement of native flower visitors by a highly abundant visitor, either
73Invasive bees and agriculture
native or alien, can decrease visit frequency by effective native pollinators
(Fig. 1C) or pollinator diversity and associated niche complementarity
(Fig. 1D) (Albrecht et al., 2012;Brittain et al., 2013a). For instance, replace-
ment of native bees and increasing transfer of self-pollen by the African honey
bee might, at least in part, explain widespread decreasing seed production in
fragmented dry forests of NW Argentina (Aizen and Feinsinger, 1994a,b). In
short, indirect costs (Fig. 1B–D) could result in the monotonic decrease of
crop yield with increasing dominance of invasive bees (Fig. 2B).
5.4 Case study 1: Bumble-bee invasion and raspberries
Studies of commercial raspberry, Rubus idaeus, in NW Patagonia provide a
detailed example of how extremely high visitation by an invasive bee can
impose direct costs on crop productivity (Morales, 2009;Sa
´ez et al.,
2014, 2017, 2018). The eruption of the Puyehue Volcano in the southern
Andes in 2011 created a natural experiment for testing the effects of invasive
B. terrestris on fruit production by commercial raspberries. A downwind gra-
dient of decreasing ash deposition caused a corresponding increase in the
average frequency of visits by B. terrestris to raspberry flowers from <5 daily
visits per flower in fields near the volcano to >150 visits in fields with little
ashfall. Surprisingly, this increased visitation did not improve the number of
carpels that developed ripe drupelets per raspberry flower. Instead, drupelet
set decreased about 40% along this gradient, so that fewer bumble-bee visits
resulted in better-quality raspberry fruits (Sa
´ez et al., 2014). The poor drupe-
let set in fields farthest from the volcano was associated with two effects of
increased visit frequency. As expected, the number of pollen grains depos-
ited on stigmas increased with visit frequency, but this effect saturated after a
few bumble-bee visits (Sa
´ez et al., 2014). The unexpected negative effect
arose from a strong positive association of the frequency of carpels with bro-
ken styles to visitation frequency, which increased throughout the gradient
from about 10% in fields close to the volcano to 90% in more distant fields
(Fig. 6A). Style breakage diminished both the chance of stigmatic pollen
deposition and, for pollinated pistils, the chance that pollen tubes reached
the ovary and fertilized the ovule (Sa
´ez et al., 2018). A mechanistic model
based on these relations predicted that just 5–10 bumble-bee visits per day—
the range of visit frequencies observed in those raspberry fields closest to the
volcano—should maximize the number of drupelets per fruit (Sa
´ez et al.,
2018). This case exemplifies the effect on plant reproduction of increasing
direct mutualism costs depicted in Fig. 2A.
74 Marcelo A. Aizen et al.
In addition to the direct costs of style breakage, increased visitation
by B. terrestris to raspberry flowers imposes an indirect cost: nectar robbery
from flower buds. Raspberry flowers produce most nectar in a single pulse
before anthesis, with little replenishment when nectar is removed following
´ez et al., 2017). When competition for nectar increases, some
bumble bees, including B. terrestris, switch from “legitimate” flower visits
to “illegitimate” robbery of nectar, thereby affecting pollination. In general,
nectar robbery damages flowers and rarely results in pollination (Irwin et al.,
2010). In particular, B. terrestris robs and damages raspberry flower buds
before they open, affecting their attractiveness and nectar availability for
“legitimate” pollinators after buds open. Sa
´ez et al. (2017) found that the
proportion of robbed buds increased from about 10 to 80% with increasing
bumble-bee visitation (Fig. 6B). Correspondingly, visitation by managed
honey bees, the second most frequent raspberry visitor in NW Patagonia,
decreased strongly with increasing visitation by bumble bees (Sa
´ez et al.,
2017). Honey bees, which are introduced but not invasive in Patagonia,
Fig. 6 Relations of interaction costs to visit rates of Bombus terrestris in fields of
raspberry (Rubus idaeus) in NW Patagonia, Argentina, including the proportions
of (A) broken styles and (B) robbed flower buds. The fitted logistic regressions are
(A) 1/(1 + e
)(z¼5.74, P<0.0001) and (B) 1/(1 + e
(Sáez et al., 2014, 2017). The images show (A) magnified (25) and microscopic
(100) pictures of undamaged (left) and damaged (right) styles and (B) pictures of
an intact and a robbed flower bud. Adapted and assembled from Sáez, A.,
Morales, C.L., Ramos, L.Y., Aizen, M.A., 2014. Extremely frequent bee visits increase pollen
deposition but reduce drupelet set in raspberry. J. Appl. Ecol. 51, 1603–1612; Sáez, A.,
Morales, C.L., Garibaldi, L.A., Aizen, M.A., 2017. Invasive bumble bees reduce nectar
availability for honey bees by robbing raspberry flower buds. Basic Appl. Ecol. 19, 1–10.
75Invasive bees and agriculture
caused less style damage to raspberry flowers than B. terrestris, but they pro-
vide similar pollination effectiveness (Sa
´ez et al., 2014). Thus, competitive
exclusion of honey bees by B. terrestris robbers would indirectly reduce
raspberry fruit production, as depicted in Fig. 1C.
5.5 Case study 2: Coffee and the African honey bee
in the Americas, revisited
Like that of B. terrestris, the invasion of the African honey bee could have
influenced crop yield in the New World. In particular, Roubik (2002)
suggested that the African honey bee enhanced crop production after it
invaded the Americas due to its high abundance and generalist foraging.
He tested this proposal by comparing mean coffee yield in tropical countries
of the Western Hemisphere before and after 1981, about the time when the
African honey bee was first observed in Central America. A paired t-test
detected statistically greater coffee production after 1981. This difference
was not detected for tropical countries of the Old World (Africa and
Asia), for which the honey-bee fauna did not change, although there was
a tendency for greater yield post-1981, as in the New World. Roubik inter-
preted this contrast as supporting the hypothesis that the African honey-bee
invasion had boosted pollination services throughout the Neotropics.
However, we question this interpretation because the arrival date of the
African honey bee differed among the New World countries involved
(Moritz et al., 2005), such that in some countries the increase in coffee yield
may have preceded the arrival of this bee.
We re-examined the hypothesis of expected benefits of honey bees for
coffee production more directly by assessing the temporal trends in yield
in individual countries with respect to the specific estimated arrival dates
of the African honey bee to each of them. A variety of patterns emerge
(Fig. 7). A minority of countries, such as Colombia, Panama, Nicaragua,
´, experienced increased average coffee yield coincident with or a
few years after the invasion of African honey bees, consistent with Roubik’s
hypothesis. Other countries, such as Bolivia, Guatemala, and Honduras,
exhibited a long-term positive trend in coffee yield that preceded the
African honey-bee invasion and was not altered by it. For these countries,
then, increases in coffee production after 1981 cannot be attributed to the
arrival of African honey bees. For a third set of countries, including Costa
Rica, Ecuador, El Salvador, and Mexico, coffee production declined
with the onset of the invasion of African honey bees or shortly thereafter.
76 Marcelo A. Aizen et al.
Fig. 7 Temporal trends of relative coffee yield from 1961 to 2017 for 17 tropical South
American and Central American countries. For each country, Δyield is the difference in
yield during year trelative to that during 1961 and transition between the white
and grey zones denotes the year of invasion by the African honey bee based on
invasion-front isoclines from Moritz et al. (2005:Fig. 2). Given the large area of Brazil,
we used 1966 rather than 1957 (i.e. when African first escaped captivity near Sao
Paulo) as the estimated average invasion year of that entire country. For each country,
the black curve describes the temporal change in mean yield (smoothing spline with a
smoothing parameter ¼0.8), whereas the green line represents the average long-term
trend (linear regression). Data from FAOSTAT, 2020. Data available at http://faostat.fao.
org/site/526/default.aspx (Accessed 25.1.2020).
This heterogeneity in response fails to support the hypothesis of wide-
ranging benefits of invasion by the African honey bee on coffee yield.
Without excluding other non-pollination explanations such as the
replacement of shade by sun coffee (Guhl, 2008), the observed variety of
effects agrees with the diversity of outcomes expected from the effects
of diverse bee densities on seed production (Section 2). Although limited
country-specific data about visitation rates and pollination precludes direct
analysis of the effects of the African honey bee on coffee yield, insights
into the nature of these effects can be gained by considering the pollination
requirements of the two most cultivated coffee species. Bothself-sterileCoffea
canephora, cultivated mostly in the Old World, and self-fertile C. arabica,
cultivated most commonly in the New World, depend on insect visitation
for high seed set and have generalist flowers that provide visitors with nectar
and pollen (Ngo et al., 2011). For example, a study of seven sets of coffee
plantations (three C. arabica, four C. canephora) in five countries found that
fruit set increased strongly with increasing visitation and diversity of wild
pollinators (Garibaldi et al., 2013). However, of the three sets visited by
A. mellifera, coffee yield varied positively with visitation rate for only the
set of plantations where honey-bee visits were in the minority. In the other
two sets, where honey-bee visits predominated, fruit set failed to show an
increase with increased visit frequency. Coffee production generally benefits
from insect pollination, but it appears that whether it benefits from
honey-bee pollination per se depends on whether A. mellifera dominates
the pollinator assemblage.
A more detailed study on C. arabica in Mexico by Badano and Vergara
(2011) is particularly revealing. Badano and Vergara found a positive effect
of overall pollinator diversity on coffee fruit set, but this effect was counter-
acted by increased numerical dominance of honey bees and consequent
reduction in the diversity of the pollinator assemblage as a whole. The
authors proposed that a diverse assemblage increases the chances of including
native pollinators that move frequently among plants, thereby carrying
outcross pollen that increases seed set (Hipo
´lito et al., 2020). Honey bees
largely restricted their foraging to individual coffee plants, transferring pollen
primarily between the flowers of the same plant. Thus, Badano and Vergara’s
results indicate indirect costs of increased honey-bee visitation to coffee
production via decreased diversity of the bee assemblage, which reduced
pollination quality. Other density-dependent direct and indirect costs
associated with active pollen harvesting (e.g. pollen theft) by honey bees
may also be involved.
78 Marcelo A. Aizen et al.
Together, this evidence suggests that honey bees may improve coffee
yield when other pollinators visit infrequently, but when alternative polli-
nators are abundant, addition of honey bees can have neutral or negative
effects. This demonstrated context dependence is inconsistent with the
conclusion that the invasion of tropical America by the African honey
bee generally improved the production of coffee or other crops.
6. The future of agriculture in a context of bee invasions
That more bee visits are always better is a prevailing paradigm in crop
pollination management (Garibaldi et al., 2020). Even though the saturation
of seed and fruit set with increasing numbers of pollinator visits is recognized
(e.g. Bell and Cresswell, 1998;Carlson, 2007), farmers often deploy more
honey-bee hives than needed “just in case” (Garibaldi et al., 2020). This
“just in case” practice is also commonly applied in the use of other agricul-
tural inputs such as fertilizers, herbicides, and pesticides (e.g. Norsworthy
et al., 2012;Schiesari et al., 2013). The conceptual framework we propose
(Section 2) and empirical evidence we present indicate that, in addition
to being unnecessary and possibly wasteful of time and resources because
lower densities of pollinators saturate plant yield capability, this practice
may actually diminish yields of pollinator-dependent crops (e.g. Brittain
et al., 2010;Carvalheiro et al., 2012).
As anthropogenic environmental disturbance increases exponentially
and climates warm, community diversity declines in all taxonomic groups
and “weedy” species, which often include successful invaders (Dawson
et al., 2011;Schlaepfer et al., 2010), increasingly dominate many landscapes
(Brook et al., 2008;Tilman and Lehman, 2001). In particular, richness
and evenness of native bee assemblages are declining globally (Arbetman
et al., 2017;IPBES, 2016;Kerr et al., 2015;Soroye et al., 2020;Zattara
and Aizen, 2019), as most pollinator species suffer but some species thrive
under habitat destruction and global warming (Aizen and Feinsinger,
2003;Bommarco et al., 2012;Herrera, 2020). For instance, B. terrestris is
increasing in relative abundance and dominating pollinator assemblages in
Scandinavia, within its native European range (Bommarco et al., 2012).
Even if overall visitation frequency by insects as a whole does not change,
indirect costs associated with the rising dominance of one or a few species
of flower visitors, such as A. mellifera or B. terrestris, will likely increase. If
these and other “weedy” generalist pollinators continue to thrive under
climate and landscape change, they may exclude specialist pollinators
79Invasive bees and agriculture
(Miller-Struttmann et al., 2015;Schweiger et al., 2010) that might pollinate
particular crops more effectively (Garibaldi et al., 2015). For example,
decreases in red clover yield in Sweden have been linked, at least in in
part, to the increasing dominance of B. terrestris and the decline of more-
specialized long-tonged bumble bees (Bommarco et al., 2012). Locally,
ill-considered pollinator management, such as overstocking a crop field
with honey-bee hives or a greenhouse with bumble-bee colonies, might
unexpectedly compromise crop yield by aggravating direct and indirect
mutualism costs (Rollin and Garibaldi, 2019;Section 2).
Current trends in land use include continuous expansion of agricultural
land cultivated with pollinator-dependent crops (Aizen et al., 2008a, 2019a).
We propose that the productivity of agricultural landscapes dominated
by pollinator-dependent crops may decline for two contrasting reasons:
too few or too many bees of just one or a few species (see also Deguines
et al., 2014). On the one hand, in extensive monocultures and other indus-
trially managed agricultural landscapes devoid of native bees and other pol-
linators, even managed bees could perform poorly and suffer high mortality
because of unbalanced diets and/or high pesticide exposure (Branchiccela
et al., 2019;Goulson et al., 2015;Mancini et al., this issue). On the other
hand, pollinators that thrive in highly disturbed habitats, such as the invasive
African honey bee, may reduce crop yield because of increasing direct
and indirect interaction costs when they are too abundant and dominant.
The results compiled and discussed in this contribution and elsewhere
(Aizen et al., 2014, 2019b) should alert governments, farmers, beekeepers,
conservationists, and other stakeholders to the detrimental consequences of
extensive trade in bees within and beyond the native ranges of the traded
species (Dicks et al., 2016;IPBES, 2016). These consequences include
negative impacts on native bee faunas and beekeeping via competition
and disease transmission, and on crop yield. Our conclusions question the
main justification for the currently booming bee trade: enhancement of
pollination services and, to a lesser extent, honey production. In addition
to sparking new bee invasions, this trade continues to augment ongoing
bee invasions such as that of B. terrestris in South America (Aizen et al.,
´rez et al., 2018). Impacts on crop yield should be thus
recognized as important considerations in risk analyses of the bee trade.
These analyses should consider the consequences of bee trade not only
for target crops, but also for other crops and native biological communities
within the countries of introduction and across international boundaries. As
the impact of B. terrestris in South America illustrates, bumble-bee trade can
80 Marcelo A. Aizen et al.
indeed enhance greenhouse tomato production in Chile (Estay, 2007),
but when they escape management these bees can impair the yield
of open-air crops, such as raspberries, and cause widespread extirpation of
native species, such as B. dahlbomii, in both Chile and neighbouring
Argentina (Aizen et al., 2019b). What is the cost of these impacts compared
to the benefits of increased tomato production? The existence of such
impacts argues for restriction of the trade in bees to the rearing and use
of native bees at a local scale.
Elimination of established invasive bees is unfortunately impractical
in most cases. Nevertheless, ongoing invasions may be slowed by stopping
the importation of non-native species instead of continuing to add “more
fuel to the fire” (Smith-Ramı
´rez et al., 2018). The agricultural impacts of
invasive bees can also be somewhat offset by enhancing conditions for native
pollinators, which generally increase crop yield when they are abundant and
diverse (Garibaldi et al., 2013), through enhancement of ecological infra-
structure at different spatial scales (Faichnie et al., this issue). At the regional
scale, this goal can be achieved by increasing landscape heterogeneity, such
as enhancing mosaics of restored natural and seminatural habitat patches
interspersed with agricultural fields cultivated with diverse crops, and by reg-
ulating agrochemical inputs (Garibaldi et al., 2014;Kova
et al., 2017). At the local scale, relevant management includes reducing
field sizes, enriching field margins with flowering plants, minimizing pesti-
cide and herbicide drifting, and providing nesting resources for bee and
non-bee pollinators (Garibaldi et al., 2014;Hobbs, 1967;Howlett et al.,
´nszki et al., 2017). The ideal result would be the
replacement of different anthropogenic inputs, including managed bees,
while promoting and supporting ecosystem services provided by naturally
occurring biodiversity through increasing ecological intensification (Garibaldi
et al., 2019;Kova
´nszki et al., 2017). In particular, increased diver-
sity of native pollinators should regulate the abundance or decrease the rel-
ative importance of super-abundant pollinators, while ensuring more stable
pollination services (Garibaldi et al., 2011b, 2013). Put simply, future
agricultural landscapes should be managed for increased biodiversity and
reduced dominance of single species, whether crop or bee.
The authors acknowledge the support of the SURPASS2 project funded under the
Newton Fund Latin America Biodiversity Programme: Biodiversity—Ecosystem Services
for Sustainable Development, grants awarded by the Natural Environment Research
81Invasive bees and agriculture
Council of Great Britain (NERC) [NE/S011870/1], the National Scientific and Technical
Research Council of Argentina (CONICET) [RD 1984/19], the Sa
˜o Paulo Research
Foundation (FAPESP) [2018/14994-1], the National Commission for Scientific and
Technological Research of Chile (CONICYT). They also acknowledge the support of
the National Fund for Scientific and Technological Research of Argentina (FONCYT)
[PICT 2015-2333, PICT 2018-2145, PICT-2018-00941], National Geographic Society
[NGS-57001R-19], Universidad Nacional de Rı
´o Negro [PI 40-B-567], the 2017–2018
Belmont Forum and BiodivERsA joint call for research proposals (under the BiodivScen
ERA-Net COFUND programme and with the funding organisations AEI, NWO,
ECCyT and NSF), and the Natural Science and Engineering Research Council of
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