![Robustness against secondary extinctions for a nested network, with no further extinctions after the loss of a specialist (a) and extinctions after the loss of a generalist (b). Functional redundancy provided by generalist species buffering against extinctions. (b) Modularity keeps disturbances within the modules of a food web, while functional redundancy can rescue the food web against extinctions. [Colour figure can be viewed at wileyonlinelibrary.com].](https://www.researchgate.net/profile/Enric-Frago/publication/347275686/figure/fig1/AS:1010295459811337@1617884616903/Robustness-against-secondary-extinctions-for-a-nested-network-with-no-further_Q320.jpg)
![Scale of the initial impact. An initial impact can lead to the loss of (a) single species, (b) functional groups or (c) parts of the network. [Colour figure can be viewed at wileyonlinelibrary.com].](https://www.researchgate.net/profile/Enric-Frago/publication/347275686/figure/fig2/AS:1010295459807235@1617884616938/Scale-of-the-initial-impact-An-initial-impact-can-lead-to-the-loss-of-a-single_Q320.jpg)
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Ecological Entomology (2020), DOI: 10.1111/een.12985
INVITEDREVIEW
Cascading extinctions as a hidden driver
of insect decline
RACHEL KEHOE,
1ENRIC FRAGO
2,3and DIRK SANDERS
1,41Centre
for Ecology & Conservation, College of Life and Environmental Sciences, University of Exeter, Penryn, Cornwall, U.K., 2CIRAD,
CBGP, Montpellier, France, 3CBGP, CIRAD, INRA, IRD, Montpellier, France and 4Environment and Sustainability Institute,
University of Exeter, Penryn, Cornwall, U.K.
Abstract. 1. The decline in insect abundance and diversity observed in many ecosys-
tems is of major concern because of the long-term consequences for ecosystem function
and stability.
2. Species in ecological communities are connected through interactions forming
complex networks. Therefore, initial extinctions can cause further species losses through
co-extinctions and extinction cascades, where single extinctions can lead to waves
of secondary extinctions. Such knock-on effects can multiply the initial impact of
disturbances, thereby largely adding to the erosion of biodiversity. However, our
knowledge of their importance for the current insect decline is hampered because
secondary extinctions are challenging to both detect and predict.
3. In this review, we bring together theory and knowledge about secondary extinctions
in the light of the main drivers of insect decline. We evaluate potential and evidence for
cascading extinction for the different drivers and identify major pathways. By providing
selected examples we discuss how habitat loss, pollution, species invasions, climate
change and overexploitation can cause cascading extinctions. We argue that habitat loss
and pollution in particular have the largest potential for such extinctions by changing
community structure, the physical environment, and community robustness.
4. Overall, cascading extinction are part of an ecosystems’ response to anthropogenic
drivers but are so far not explicitly measured in their contribution when evaluating
biodiversity loss. This knowledge is necessary to predict biodiversity loss and nd
strategies to buffer against the devastating long-term impact of habitat loss, pollution,
species invasions, and climate change.
Key words. apparent competition, climate change, co-extinction, functional extinction,
habitat loss, indirect effects, pollution, secondary extinction, species invasions.
Introduction
Many studies report marked declines in insect abundance and
diversity across ecosystems and taxa (Hallmann et al., 2017;
Leather, 2018; Habel et al., 2019; Montgomery et al., 2020).
However, while this negative trend holds for terrestrial insects,
the authors of a recent meta-analysis found much variation
Correspondence: Dirk Sanders, Centre for Ecology & Conservation,
College of Life and Environmental Sciences, University of Exeter,
Penryn, Cornwall TR10 9FE, U.K. Environment and Sustainability
Institute, University of Exeter, Penryn, Cornwall TR10 9FE, U.K.
E-mail: d.sanders@exeter.ac.uk
in the included datasets when observing different taxa and
locations (van Klink et al., 2020). Overall, there is clear evi-
dence that insects are declining and have done so for years in
many habitats around the world, which is driven by changes in
land-use, pollution, biological interactions (mostly invasions),
and climate change (Wilson & Fox, 2020.; Seibold et al., 2019;
Sánchez-Bayo & Wyckhuys, 2019; Cardoso et al., 2020; Wag-
ner, 2020).
This overall negative trend for many insect populations
has severe consequences for ecosystem stability and function
(Soliveres et al., 2016; Fanin et al., 2018). First, there is a
positive relationship between insect diversity and many of the
functions they provide (e.g., Srivastava & Bell, 2009; Cardinale
© 2020 The Royal Entomological Society 1
2Rachel Kehoe et al.
et al., 2012). Second, the interconnectedness of species in eco-
logical communities allows an impact on single species to be
transmitted to the rest of the community. Initial extinctions can
therefore trigger further extinctions passed on through direct and
indirect interactions that subsequently start a process destabil-
ising whole ecological communities. Of particular concern are
cascading extinctions, where this initial impact causes waves
of species extinctions, thereby eroding biodiversity (Säterberg
et al., 2013; Sanders et al., 2018a). This can lead to a reduction in
functional redundancy (Sanders et al., 2018a; Biggs et al., 2020)
and ultimately to an increased vulnerability to future distur-
bances.
In this review, we rst discuss the theoretical background
of secondary extinctions, which can be triggered by the loss
or decline of resources, consumers, competitors, mutualists, or
ecosystem engineers. We then explore the potential of such
cascading extinction in the light of the main drivers of bio-
diversity decline by using selected examples from the litera-
ture. Ecological theory predicts that secondary or cascading
extinctions are a threat responsible for many regional and local
extinctions (Borrvall & Ebenman, 2006; Säterberg et al., 2013;
Brodie et al., 2014). However, empirical evidence is hard to
obtain because we still lack a more mechanistic understanding
of how the initial disturbance is transmitted through interac-
tions in ecological communities leading to secondary extinc-
tions. Further it is often difcult to distinguish between cascad-
ing extinctions and background noise imposed by the constant
environmental change caused by human activities. Therefore,
even if theory (e.g., Thébault et al., 2007; Brodie et al., 2014)
suggests that cascading extinctions are very important for our
understanding of biodiversity loss, few experimental studies
have explored this question (e.g., Donohue et al., 2017; Sanders
et al., 2018a). In this paper, we argue that we still lack empir-
ical knowledge to draw general conclusions, particularly stud-
ies of secondary extinctions that include whole community
dynamics. Here we focus on insects, which are crucial parts
of the wider community of different taxa, but a broader com-
munity approach is necessary to understand the role of cascad-
ing extinctions in driving their decline. We acknowledge the
importance of considering insect links to other animal groups,
and of experimental data coming from non-insect groups to
draw general conclusions on the mechanisms behind insect
extinctions.
We bring together theory about the extinction cascades and
current evidence of how major anthropogenic drivers lead to
insect decline by presenting key examples and linking them
to ecological mechanisms. This will help to evaluate the role
of run-away extinctions and nd strategies to predict which
ecological communities may be most at risk. Specically, we
rst identify mechanisms for cascading extinctions and discuss
pathways for the main drivers of insect decline to initiate such
cascades. We then summarise the current evidence for cascades
of secondary extinctions as hidden drivers for insect decline.
Finally, we discuss the uncertainties and suggest ways for future
research that can solve these uncertainties and increase our
ability to predict the strength of extinction cascades, and when
they are most likely to occur.
Predicting secondary extinctions
In this section, we review basic theory about secondary extinc-
tions with initial disturbances leading to varying consequences
for community structure and stability; we apply this in the next
section to the impact of drivers on insect populations. The mag-
nitude of the follow-on effects of the initial loss depends largely
on the function or role of the species going extinct and the ability
of the rest of the ecological community to compensate or buffer
for that loss.
We can divide secondary extinctions into two main cate-
gories: follow-on single extinction events (e.g., co-extinctions),
and cascading extinctions leading to further extinction events
along the way as the impact is passed on to more and more
species. Secondary extinctions can be triggered by the extinc-
tion of a species or a signicant decline in abundance. The latter
is potentially widespread because a species’ role in a commu-
nity depends on its population density. The function can be lost
even when the species is still present, because it has declined
below a certain threshold density: this impact is called ‘func-
tional extinction’ (Säterberg et al., 2013; Sanders et al., 2018a).
Many studies report a substantial decline in insect abundance
and biomass with negative trends for many populations across
the world (Hallmann et al., 2017; Seibold et al., 2019; Wag-
ner, 2020), therefore, we can indeed expect this decline to be
a major driver for secondary extinctions. Secondary extinctions
are likely occurring at smaller landscape scales because at this
scale populations interact more tightly with each other with sig-
nicant consequences for ecosystem functions within local com-
munities. However, declines and extinctions of local population
also contribute to global extinctions and increase the overall
extinction risk for species (Blaustein et al., 1994).
Co-extinctions and network transmitted extinctions
The majority of studies consider secondary extinctions to be
co-extinctions (Brodie et al., 2014; Veron et al., 2018; Cardoso
et al., 2020): the direct dependence of one species on another
leads to its demise, as for example when a parasite or a spe-
cialist predator goes extinct because their host or resource have
disappeared (Fig. 1a). Co-extinctions can involve many differ-
ent types of interactions, such as pollination, seed dispersal,
symbiosis, mutualism, parasitism, predation, and non-trophic
interactions and are most likely to affect resource or habitat
specialists. For instance, a low population size of its hosts, the
endangered black (Diceros bicornis (Linnaeus, 1758)) and white
(Ceratotherium simum (Burchell, 1817) rhinoceroses, put the
stomach bot y Gyrostigma rhinocerontis (Owen, 1830) which
is a specialist parasite (Colwell et al., 2009), in an even more
endangered situation. We can expect the effect to be immedi-
ate for intimate interactions such as symbiosis or parasitism,
or the absolute necessity of an ecosystem engineer such as a
tree providing the habitat for an insect. As a possible common
example of resource driven co-extinctions, the extinction of but-
teries on a tropical island in Singapore was a result of the loss
of their host plants (Koh et al., 2004). The impact on the depend-
ing partner can be delayed in long-lived species if the organism
has the ability to survive on its own but relies on the extinct
© 2020 The Royal Entomological Society, Ecological Entomology, doi: 10.1111/een.12985
Insect decline and extinction cascades 3
(a) (b)
Fig. 1. Secondary extinctions. (a) Co-extinctions after the initial loss of a resource, (b) network transmitted extinctions driven by changes in interaction
strength and indirect interactions. Red nodes (with an x) in the food web are going initially extinct with the orange nodes as secondary extinctions. The
arrows indicate the transmission of the initial impact. [Colour gure can be viewed at wileyonlinelibrary.com].
species for reproduction. For example, the extinction of larger
vertebrates on islands poses a problem to seed dispersal and off-
spring recruitment for larger trees (Wang et al., 2007). A recent
study done on Réunion island used lava chronosequences that
spanned 300 years and revealed the strong impact of frugivore
extinctions on forest community structure (Albert et al., 2020).
The lower functional redundancy of island ecosystems sub-
stantially increases the risk of co-extinctions (Kaiser-Bunbury
et al., 2010).
Species in ecosystems form complex networks of interac-
tions; this means the initial loss of a species can be transmit-
ted along the interactions in networks and cause further extinc-
tions, potentially to parts of the network found many links
away (Fig. 1b). There is evidence that species that are tar-
geted by harvesting are not the ones that go extinct in the rst
instance (Sanders et al., 2015). Concern should therefore be
raised about the potential extinction of other indirectly con-
nected species in the network. This impact can be transmit-
ted through direct and indirect interactions. Indirect extinctions
occur when the effect between two species is passed through
at least a third one. From a large body of research in commu-
nity ecology, we know of the importance of indirect interactions
such as apparent competition, trophic cascades, apparent mutu-
alism and higher order interactions for dynamics, stability, and
functions (Wootton, 1994; Morris et al., 2004). Trophic cas-
cades are a powerful example of indirect interactions, where
diminished top-down control leads to a marked negative impact
on primary producers (Schmitz, 2003; Estes et al., 2011). The
removal of otters through hunting, for instance, led to the release
of sea urchins from top-down control and the destruction of
kelp forest, with a complete shift from diverse kelp forests to
barren sea oor communities (Estes & Palmisano, 1974). The
long, shared ecological and evolutionary histories of island tor-
toises and plant communities has shaped many plant-tortoise
interactions, many of which have since been lost as a result of
tortoise decline or extinction (Sobral-Souza et al., 2017). For
example, ‘tortoise turf’ is a plant community of endemic grass,
herb, and sedge species, which is engineered by continuous tor-
toise grazing and trampling. It is thought to have been common
on islands throughout the Indian Ocean before tortoises went
extinct; it is now restricted to Aldabra (Merton et al., 1976;
Cheke & Hume, 2010). This demonstrates that predator or con-
sumer extinctions can have far-reaching consequences for whole
ecosystems. Further, the loss of diversity or complexity in a com-
munity can indirectly lead to further extinctions through the loss
of associational resistance. For example, the reduced diversity
of non-hosts in a community has been shown to increase para-
sitism of an insect host, leading to its overexploitation (Kehoe
et al., 2016).
A bottom-up extinction cascade is driven by the loss of a
resource, which can be either food, habitat, or a condition that
was provided by the presence or activity of another organism
(e.g. a tree as a habitat for bats and many insects). In Yel-
lowstone, the invasive lake trout Salvelinus naymaycush (Wal-
baum, 1792) reduced densities of the native cutthroat trout,
Oncorhynchus clarkii bouvieri (Richardson), triggering large
changes in aquatic arthropod communities which then triggered
a bottom-up effect on terrestrial communities of large predators
including mammals and birds (Koel et al., 2005, 2020; Tron-
stad et al., 2010). This effect can likely cascade to insects as
shown in Knight et al. (2005) who found that sh presence
in ponds altered populations of aquatic insects and indirectly
pollinators and pollination of terrestrial plants. The combined
effect of top-down and bottom-up cascading effects can trans-
late into horizontal extinction cascades as shown in (Sanders
et al., 2015, 2018a). Predators can be necessary for the coexis-
tence of lower trophic levels, for example, if predators regulate
the density of the most competitive prey species. A predator
extinction can thus lead to dominance shifts in prey and the
extinctions of other predators that depend on them. Horizon-
tal extinctions cascades are predicted to happen when consumer
guilds are specialized to a certain degree, which means their
function as predators is necessary. Research has shown that the
system can be rescued by generalist predators providing func-
tional redundancy and loosely linking different food web mod-
ules (Sanders et al., 2018a).
Most research on indirect extinctions has focused on effects
driven by changes in the density of interacting species. Species
declines, however, can also be transmitted via behavioural
changes (also known as trait- mediated), but their impact on sec-
ondary extinctions is so far little studied. Trait-mediated indirect
effects are currently acknowledged as important in ecological
communities (Peacor & Werner, 2001; Schmitz et al., 2004).
For example, predator avoidance behaviours may allow species
persistence by reducing the effectiveness of highly competitive
competitors, a prediction that has been demonstrated by recent
© 2020 The Royal Entomological Society, Ecological Entomology, doi: 10.1111/een.12985
4Rachel Kehoe et al.
(a) (d)
(e)
(b)
(c)
Fig. 2. Robustness against secondary extinctions for a nested network, with no further extinctions after the loss of a specialist (a) and extinctions after
the loss of a generalist (b). Functional redundancy provided by generalist species buffering against extinctions. (b) Modularity keeps disturbances
within the modules of a food web, while functional redundancy can rescue the food web against extinctions. [Colour gure can be viewed at
wileyonlinelibrary.com].
models showing that anti-predator avoidance can allow commu-
nity stability (Sommers & Chesson, 2019). Fear-induced avoid-
ance behaviour has been shown to promote species coexistence
in many communities (Leibold, 1991; Werner, 1992). Research
shows that fear of large carnivores reduces racoon activity and
has a positive effect on racoon prey including insects (Suraci
et al., 2016). Loss of these top predators can thus trigger extinc-
tion cascades through the loss of trait-mediated indirect interac-
tions, but more research is needed in more complex scenarios
that for example combine different types of interactions in mul-
tilayer networks.
The role of network structure in transmitting the effect
of species extinctions
The most insidious extinction cascades are those that erode
the necessary structural network components for community sta-
bility (resistance, resilience, or robustness). Such a structural
breakdown leaves the community vulnerable and is often asso-
ciated with regime shifts and erosion of biodiversity (Hughes
et al., 2007; Carpenter et al., 2011). For example, nestedness
has been shown to increase robustness in mutualistic networks
such as plant-pollinator systems (Thébault & Fontaine, 2010). A
nested network is characterised by a core of generalist species
that link the whole network while any specialist is using only
links that are already provided by generalists. This means the
system is robust against specialist extinctions (Fig. 2a) but if a
generalist from those communities is lost (Fig. 2b), this will have
a marked effect on other species’ persistence in the community
(Bastolla et al., 2009; Thébault & Fontaine, 2010). Antagonistic
networks such as food webs on the other hand gain increased sta-
bility through modularity (Fig. 2d, Thébault & Fontaine, 2010)
where species interact in subnetworks that are not or only weakly
linked. Theory predicts that this structural feature keeps a dis-
turbance with a module rather than letting it spread through
the whole network (Krause et al., 2003). Other studies, how-
ever, suggest that connectance (the proportion of realised links
in a network) and functional redundancy (redundant links in a
network, see Fig. 2c,d) are important structural components to
buffer against secondary extinctions in antagonistic networks
(Dunne et al., 2002; van Altena et al., 2016). Possibly there
is truth in both: modules that are loosely linked by generalist
predators (Fig. 2e) with a certain level of functional redundancy
are likely to be most robust to cascades of secondary extinc-
tions as shown in (Sanders et al., 2018a). This may explain
the importance of generalist top predators, often called key-
stone predators, in maintaining stability because they prey on the
most abundant prey and thereby enhance the survival of inferior
competitors often allowing their coexistence within the commu-
nity at different trophic levels (Paine, 1995). The loss of such
a keystone predator can have far reaching consequences as an
important structural component is lost (Estes et al., 2011). This
means that as communities become simpler with biodiversity
loss, networks become more vulnerable to cascading extinctions
(Borrvall et al., 2000; Sanders et al., 2018a).
© 2020 The Royal Entomological Society, Ecological Entomology, doi: 10.1111/een.12985
Insect decline and extinction cascades 5
(a) (b) (c)
Fig. 3. Scale of the initial impact. An initial impact can lead to the loss of (a) single species, (b) functional groups or (c) parts of the network. [Colour
gure can be viewed at wileyonlinelibrary.com].
Scale of the initial impact
To determine how important secondary extinctions are, and
which ecological communities are more affected, we need
knowledge on how often the initial disturbance hits and how
different drivers of species extinctions interact. This initial
impact can affect a single species (e.g., overexploitation) or the
whole community (e.g., habitat loss) (Fig. 3a). It is difcult to
predict which outcome is more likely to lead to more severe
cascading extinctions because this very much depends on the
role of the lost species. For example, predators might be more
vulnerable than other trophic levels (Fig. 3b, Purvis et al., 2000)
and if the coexistence at the lower trophic levels depends on
top-down control we can expect secondary extinctions after
the loss of the predator guild. As discussed above, modularity
increases robustness in food webs, so that the extinction of
a whole sub-network can potentially be contained within that
part of the network with very little follow on extinction events.
This, again, depends on the function of any bridging species
to the remaining network. The loss of beavers and their dam
building activity (Wright et al., 2002) will be most severe for the
aquatic food web with weaker knock on-effects on the terrestrial
community (Fig. 3c). If the ow of energy into the terrestrial
food web is substantial (as shown in the Yellowstone example
above), however, this will lead to more important shifts in the
community structure.
At the temporal level, in some instances the impact can be
immediate, such as for many co-extinctions. These are therefore
easier to detect and predict from observational data. Network
transmitted or functional extinction events will be much harder
to uncover because they will be delayed through population
dynamics and feedbacks (Sanders et al., 2015). Finally, if the
disturbance has already led to irreversible changes in ecological
communities, for example, by driving populations to a very low
densities, or through the loss of mutualistic relationships, we can
expect to see extinctions happening in the future, a phenomenon
known as an extinction debt.
Drivers of insect decline and cascading extinctions
Five main drivers have been made responsible for the decline
in insect abundance and diversity (Habel et al., 2019;
Sánchez-Bayo & Wyckhuys, 2019; Cardoso et al., 2020;
Fiza Fatima et al., 2020; Montgomery et al., 2020; van Klink
et al., 2020; Wagner, 2020): (I) habitat loss and fragmentation
associated with housing development and agriculture, tim-
ber, and livestock production, (II) pollution through spill of
toxic chemicals into the environment and direct application of
pesticides, as well as light and noise pollution: (III) invasive
species, (IV) climate change and (V) overexploitation. There
is a high degree of linkage between some of these drivers,
with for example habitat loss driven itself by climate change.
Additionally, individual drivers can interact with each other,
with one increasing the severity of another. Climate change, for
example, can alter the infection rate of Nosemosis in honeybee,
Apis mellifera Linnaeus 1758 (Martín-Hernández et al., 2009),
while other factors such as pesticide load or decreased resource
availability reduce the health of a population and leave the popu-
lation more vulnerable to pests or pathogens (Potts et al., 2010).
In order to understand, predict, and prevent insect losses from
ecosystems, we need to identify the details of the pathways that
lead to this loss. Here we look at the potential of cascading
extinctions associated with each driver and their mechanisms
and present a non-exhaustive list of examples.
Habitat loss and degradation
A recent review found that about 50% of articles research-
ing insect decline, report habitat changes as the main driver
(Sánchez-Bayo & Wyckhuys, 2019). This highlights its impor-
tance in diminishing insect populations through the expansion
and intensication of agriculture, forestry, livestock production,
urbanisation, and mining (Brook et al., 2003; Habel et al., 2019;
Mammola et al., 2019). This happened especially when agri-
cultural practices shifted in many countries from traditional,
low-input farming style to intensive, industrial-scale production
(Bambaradeniya & Amerasinghe, 2004; Ollerton et al., 2014).
Although habitat loss can directly kill insects when habitats are
transformed, extinction events are likely to be caused by indirect
effects such as the loss of resources and alteration of the physical
environment.
Agriculture and deforestation substantially change natu-
ral plant communities, with those communities becoming
homogenised and simplied, the removal of long-lived
plants, and the timing of owering being changed (Thomas &
Kevan, 1993). Agriculture often turns a natural environment into
a habitat dominated by a select few plants, while urbanisation,
intensive livestock production and mining can be even more
© 2020 The Royal Entomological Society, Ecological Entomology, doi: 10.1111/een.12985
6Rachel Kehoe et al.
(a) (b)
(c) (d)
Fig. 4. Drivers and cascading extinctions. Arrows with continuous lines depict direct interactions. Indirect effects of drivers on insect extinctions are
also shown as dashed lines. The initial impact through drivers is shown by a red circle. Higher-trophic levels suffer from habitat loss as bottom-up
effects are magnied along food chains (a1). Habitat degradation can increase the density of insect natural enemies and trigger insect extinctions (a2).
Highly fertilized agricultural habitats increase pest densities with negative indirect effect on other insects through apparent competition (a3). The impact
of habitat loss depends on insect functional traits including dispersal capabilities, longevity and specialisation (a4). Pollution in the form of herbicides
and insecticides bioaccumulates along food chains (b1). These products can kill non-target organisms and trigger declines of pollinating insects thus
altering plant-pollinator networks and plant communities (b2). Anthelmintic substances used to treat worm infestation in livestock alter dung beetle
communities (b3). Since some of these beetles are plant pollinators, this can also alter plant-pollinator networks (b2). (c) Invasive species can trigger
extinctions of local insects indirectly through shared natural enemies (i.e. apparent competition, c1), or by altering local plant communities and trigger
bottom-up extinction cascades (c3). Invasive predators can become top-predators via intraguild predation and release certain populations of prey from
top-down pressure thus increasing pressure on others (c2). (d) Climate change alters the synchrony between plants and herbivores. Some omnivorous
herbivores (here mice) also predate on insects thus controlling their populations. This equilibrium can be altered, and pest outbreaks triggered with
negative effects on non-pest species via resource competition (d1). Some insects rely on mutualistic symbioses (represented here as a red bacterium)
to obtain protection from natural enemies. These symbionts are often susceptible to increased temperatures so that global warming can render these
insects unprotected and susceptible to demise due to top-down pressure (d2). [Colour gure can be viewed at wileyonlinelibrary.com].
extreme in removing the majority of plants from habitats. Both
processes can lead to co-extinctions of many plant-associated
insects. For example, the clearance of vegetation in Singapore
caused the loss of 208 plant species upon which specialist but-
tery species rely, with at least 56 buttery species becoming
co-extinct (Koh et al., 2004). This has likely resulted in a cas-
cade of extinctions as other species such as other herbivores and
predators/parasitoids associated with these food web modules
will have also lost their resources (Fig. 4a1). We can expect
resource-driven extinction cascades to be very common that
affect many species. Vegetation structure is providing necessary
habitat structure needed as shelter and refuge for many species
while maintaining a specic microclimate. All plants are acting
therefore essentially as ecosystem engineers (Jones et al., 1996;
Sanders et al., 2014), with species and interactions between
species depending on the state of this physical environment.
In addition to providing a food resource, trees, for example,
provide numerous abiotic requirements for insect survival such
as both substrate for nests and webs (Santos & Gobbi, 1998).
This includes plant bre as an important material for nest
building (Rodrigues & Machado, 1982), shading, and as such
protection from direct sunlight and high temperatures, as well
as desiccation, and camouage against predators. Deforestation
and logging remove these structures, which will indirectly
cause the loss of species and change the strength of interactions
between species (e.g., Chase, 1996). Declines in moths, for
example, are linked to the availability of their overwintering
host plants (Mattila et al., 2006; Fox, 2013), specialist ground
beetles depend on hedgerows and trees (Brooks et al., 2012) and
overall species diversity declines with decreasing vegetation
© 2020 The Royal Entomological Society, Ecological Entomology, doi: 10.1111/een.12985
Insect decline and extinction cascades 7
complexity in many groups of organisms (Tews et al., 2004).
Other ecosystem engineers can have a similarly marked role;
with ants for example increasing grazing quality for cattle
(Li et al., 2018), and shifting the balance of top-down and
bottom-up effects (Zhong et al., 2017). The removal of ants,
which are very sensitive to land use, can lead to reduced abun-
dance of other insects and impact the diversity at higher trophic
levels (Sanders & Veen, 2011).
Habitat change impacts the way species interact through the
loss of refuges or by increasing the ability of predators to use the
habitat or by limiting primary producer biomass; this can then
lead to extinctions and important shifts in community structure.
For example, urbanisation has been shown to promote the
density of some avian predators (Evans et al., 2009) through the
provision of nesting sites which can ultimately increase pressure
and trigger top-down extinctions of insect prey (Fig. 4a2).
Urbanisation has also been shown to signicantly change the
way species interact by reducing the strength of a trophic
cascade (Turrini et al., 2016). Uncovering these shifts needs
well-designed experiments, with the results then scaled up to the
habitat level. Habitat loss can shift the balance between species
that interact through apparent competition (Fig. 4a3). This may
be common in temperate agricultural systems because they are
often irrigated and fertilised, making them highly productive
and therefore hosting large densities of insect pests (Garratt
et al., 2011; Butler et al., 2012). These increased pest densities
can lead to higher densities of generalist predators that can spill
over to nearby natural habitats diminishing other insects (Rand
& Louda, 2006).
Together with habitat loss, changes in habitat conguration
has been described as ‘the single greatest threat to biological
diversity’ (Noss, 1991). Insect diversity tends to decline in small
forest fragments after the surrounding habitat is lost (Kruess
& Tscharntke, 2000; Jennings & Tallamy, 2006). As predicted
by island biogeography theory (MacArthur & Wilson, 2001),
in such small fragments even if the vital space to persist
is available, crucial elements needed for species to persist
like particular plant resources, refugees, or other species (i.e.
plants, other arthropods or microbes) may be missing. Smaller
habitats are only capable of supporting small populations, which
are usually linked to higher extinction risks. This unstable
system is vulnerable to cascades of extinctions as the essential
connectivity between species is eroded. Changes in habitat
conguration, like increased fragmentation, may also exasperate
the effects of climate change as poor connectivity between
habitats constrains range shifts (Platts et al., 2019), particularly
in species with low dispersal capabilities (Fig. 4a4). Food-web
theory and empirical evidence suggest that higher trophic levels
are more susceptible to disturbances than lower trophic levels
(Purvis et al., 2000; Binzer et al., 2011), and it is therefore likely
that they suffer disproportionately with the potential of causing
extinction cascades (Estes et al., 2011) (Fig. 4a1).
The co-extinctions due to a loss of resources are therefore
likely to be fast, because without their resources herbivores
and their associated predators will starve. However, cascading
extinctions due to loss of physical environment may happen at
a slower pace. As found in butteries, longer-lived species may
persist in degraded habitats for longer than short-lived ones but
are unlikely to persist in the long term (Krauss et al., 2010).
These species can thus represent an indirect ‘extinction debt’
due to habitat loss and degradation through agricultural inten-
sication approximately 40 years prior, indicating that further
cascading extinctions are likely to continue. Relative to general-
ists, habitat or resource specialists are more likely to suffer from
co-extinctions and cascading extinctions (Fig. 4a4). In agricul-
tural and forestry lands, or remnants of natural habitats derived
from urban settlements, the new habitats created are less likely
to full the requirements that specialists need to survive, for
instance, because the few plants they rely on are less likely to
remain (Praz et al., 2008). These alterations leading to simpler
communities may leave fewer more common, generalist species
giving a higher network connectance, which can increase robust-
ness (Dunne et al., 2002). However, the functionality of the net-
work is likely to be altered, which can itself lead to further loss.
Pollution
Sánchez-Bayo and Wyckhuys (2019) report that a quarter of
all studies about insect decline in their review mention pollution
as the cause. They subcategorised pollution into the application
of pesticides (12.6%), fertilisers (10.1%), and the spill of other
pollutants (3.1%) such as industrial, noise, and light pollution
into the environment.
Insecticides have a very high potential to trigger secondary
extinctions by causing the functional extinction of many insect
species directly and by passing on the effect through popula-
tion dynamic effects and bioaccumulation through food webs
(Fig. 4b1). Insecticides for both agricultural and human use
cause mortality of both the target and non-target species by
direct intoxication as well as reducing their health and fecun-
dity. These effects may result in shifts in the abundance and
diversity of many insects, such as the use of neonicotinoid insec-
ticides reducing wild bee density, solitary bee nesting, bum-
blebee colony growth, and reproduction (Rundlöf et al., 2015)
(Fig. 4b2), but also killing natural enemies as they are trans-
ferred from the plant via honeydew (Calvo-Agudo et al., 2019).
Despite this evidence, we still lack knowledge about the
responses of many other groups, though see Córdoba-Aguilar
and Rocha-Ortega (2019) for evidence of wastewater pollutants
reducing the tness of damselies. Bioaccumulation can lead
to high levels of insecticides in higher trophic levels (Hayes &
Hansen, 2017). The bioaccumulation of the malaria preventa-
tive insecticide DDT, for example, increases in concentration
with higher trophic levels such as in carnivorous coccinellid
beetles and arachnids (Rudd et al., 1981), potentially leading
to cascading extinctions. Decomposers, such as dung beetles,
tend to suffer from the use of anthelmintic substances in treat-
ing worm infestation in livestock (Verdú et al., 2018) (Fig. 4b3).
This indirect loss of dung beetles will disrupt the ecological
services which they provide such as nutrient cycling, bioturba-
tion, plant growth enhancement, secondary seed dispersal, and
parasite control (Nichols et al., 2008). A number of dung bee-
tles are also (often obligate) pollinators of decay-scented ow-
ers (Nichols et al., 2008) with the loss of these beetles directly
causing co-extinctions of plants (Fig. 4b2). Herbicides damage
© 2020 The Royal Entomological Society, Ecological Entomology, doi: 10.1111/een.12985
8Rachel Kehoe et al.
or kill plants in areas which they are not wanted, such as in
with other crops or in urban environments to increase aesthetics.
Hawes et al. (2003) showed that herbicide application reduced
weed vegetation, which linked directly to reduced pollinator
abundance, and herbivore presence, and consequently reduced
abundance of associated predators and parasitoids. Detritivore
abundance was also seen to reduce due to the decreased quan-
tity of dead plant material associated with herbicide use. The
use of herbicides has also been shown to reduce body condi-
tion in some dung beetles (Villada-Bedoya et al., 2019), with
the consequent loss of ecosystem services that they provide. As
discussed above in the habitat loss section, the loss of plants will
cause co-extinctions of both species requiring them as a food or
depend on their ability to provide habitat structure and certain
abiotic conditions. Osborne et al. (1991) suggest that the reduc-
tion of pollination through the loss of bees will likely result in
the local extinction of species pollinated by these bees. These
co-extinctions can turn into a network transmitted cascade as the
reduced ower, seed, and fruit production will result in a decline
in pollinators, frugivorous, and granivorous species, as well as
those that depend on the plants for reproduction or habitat.
In addition to chemical pollution, articial light at night has
been shown to alter physiological and behavioural traits such
as cognition, offspring number, hormone levels, and survival
of species across many taxa (Sanders et al., 2020) as well
as attracting insects in particular (e.g. Poiani et al., 2005).
This in turn has important implications for their overall tness
(Shimoda & Honda, 2013) and is linked to global insect decline
(Kruskal, 2018). Therefore, we can expect knock on effects on
other species in ecological communities. For example, articial
light at night can change interactions within ecological networks
such as altering predation rates (Minnaar et al., 2015; Sanders
& Gaston, 2018; Sanders et al., 2018b; Kehoe et al., 2020)
and pollination, impacting entire networks (Knop et al., 2017).
These changes to the structure of networks and can cause so far
undetected cascading extinctions.
Invasive species
Only 2% of articles (Sánchez-Bayo & Wyckhuys, 2019)
describe insect decline through invasive species. This area may
be understudied, or the impact on insects is less common than for
other drivers, however, for example, due to their large population
sizes, aggressive behaviour, and generalist feeding, invasive
ant species can have powerful impacts on native communities
(Human & Gordon, 1997; Sanders et al., 2003). Invasive
species can directly impact others, such as by outcompeting
natives, causing direct extinctions or bringing novel parasites,
diseases or pathogens. Indirect effects include changing native
community structure for example by outcompeting native host
plants (Tallamy et al., in press) or causing changes in behaviour,
such as altering feeding through predator avoidance or increased
movement, causing increased energy expenditure.
Many examples suggest that invasive species can have indirect
effects on insect populations via apparent competition, i.e.
via shared natural enemies (Fig. 4c1). The outbreaking moth
Lymantria dispar (Linnaeus, 1758) is native from Europe and
invaded America in 1868 or 1969 (Elkinton & Liebhold, 1990).
To eradicate this pest, the generalist parasitoid y Compsilura
concinnata (Meigen, 1824) was introduced. This natural enemy
tracked L. dispar during its invasion across the continent, but
it did not control its populations. This generalist parasitoid,
however, became very abundant in areas with large densities
of the invasive moth, often spilling over from its primary
host and attacking local butteries like the Papilio canadensis
(Rothschild and Jordan, 1906) (Redman & Scriber, 2000).
C. coccniata is a polyphagous and multivoltine parasitoid
that completes one generation on the invasive L. dispar and
subsequently attacks other native lepidopterans in late summer.
These species include the emblematic giant silk moths, which
are suffering vastly reduced populations as they get more than
80% parasitism (Elkinton & Boettner, 2004).
Apparent competition and their consequences for insect
extinctions can also be inferred by building quantitative trophic
webs. Henneman and Memmott (2001) demonstrated that exotic
plants, insects, and insect natural enemies can have profound
effects on native communities via indirect effects transmit-
ted through trophic networks. The authors of this study sug-
gested that some alien herbivores introduced to control inva-
sive plants like the bramble, Rubus argutus Link 1822, ginger
Hedychium gardnerianum (Roscoe), and Acacia melanoxylon
(Brown), could increase densities of certain parasitoids capable
of attacking native host caterpillars, potentially triggering their
demise.
As mentioned earlier, top predators and other species found
at the higher levels of food chains are particularly vulnerable to
extinction (Post, 2002). In many cases, however, invasive preda-
tors can establish novel higher-order interactions and become
keystone predators in novel habitats driving indirect extinctions
of local prey (Fig. 4c2). A recent study experimentally intro-
duced curly-tailed lizards, Leiocephalus carinatus Gray 1827,
in small Caribbean islands, where brown anole lizards, Anolis
sagrei (Dumeril and Bribron, 1837), were usually top-predators
(Pringle et al., 2019). The invasive species fed on local lizards as
an intraguild predator and displaced the local lizard from its top
position in the trophic web. This new predation risk changed the
behaviour of the local predator ultimately reducing consumption
of cockroaches and ants and increasing consumption of beetles.
This example illustrates the importance of trophic web studies to
understand insect extinctions as novel scenarios can render infe-
rior competitors below survival thresholds. Invasive ants have
been shown to impact many species, thereby changing commu-
nity structures and ecological functions (Gotelli & Arnett, 2000;
Sanders et al., 2003). Adding a new aggressive generalist preda-
tor to a community will have far-reaching consequences for prey
species and other predators causing top-down and horizontal
extinction cascades.
Invasive plants and their associated herbivores can also trigger
indirect species extinctions up in the trophic web (Fig. 4c3).
Probably one of the best-documented cases of an invasive
insect triggering direct extinctions is that of the harlequin
ladybird, Harmonia axyridis (Pallas, 1773), in Britain, Wales
and Scotland (Roy & Brown, 2015). This beetle predator feeds
mostly on aphids, but also via intraguild predation on other
ladybirds. In America, the negative impact of this invasive insect
© 2020 The Royal Entomological Society, Ecological Entomology, doi: 10.1111/een.12985
Insect decline and extinction cascades 9
has increased via indirect effects driven by the invasive plant,
common buckthorn Rhamnus cathartica (Linnaeus). This plant
has served as an important host for the invasive soybean aphid,
Aphis glycines (Matsumura), which has become a major prey
for H. axiridis, allowing this latter species to spread and reach
high local densities (Heimpel et al., 2010). Similar bottom-up
effects have been found due to the invasion of the emerald
ash borer, Agrilus planipennis Fairmaire 1888, in America.
This beetle is causing massive mortality on trees in the genus
Fraxinus. This tree genus is very diverse in the Americas, and
many specialist insects thrive on them. A recent literature review
has identied at least 45 species (Gandhi & Herms, 2010),
particularly Lepidopterans that depend on these trees and that
are at indirect risk of extinction due to the invasive beetle
(Wagner, 2007).
Changes brought about by invasive species in their local
habitats can also ripple through distant habitats via indirect
interactions. In the Yellowstone example mentioned above,
for example, the invasive lake trout, Salvelinus namaycush
(Walbaum, 1792), is a top predator that has triggered the decline
of the local Yellowstone cutthroat trout, Oncorhynchus clarkii
(Richardson, 1836), and caused marked community changes
for water arthropods (Koel et al., 2020). These changes have
also cascaded into terrestrial consumers like bears, which in the
absence of O. clarkii are increasingly predating on elks, Cervus
elaphus Linnaeus,1758 (Koel et al., 2005). This study did not
investigate effects on arthropod communities, but water to land
effects are likely in this invasion scenario, given the strength of
the cascading interactions found.
Climate change
Global warming is changing plant and animal phenology.
This can lead to co-extinctions if these changes affect different
trophic levels thus leading to phenological mismatches between
insects and plants (van Asch & Visser, 2006), and between her-
bivores and natural enemies (Schreven et al., 2017). In tritrophic
systems, increased warming can also trigger indirect bottom-up
extinctions. Current projections of increased spring warming,
for example, are predicted to increase mismatches between oak
but burst and winter moth, Operophtera brumata (Linnaeus,
1758), caterpillars, indirectly affecting predatory birds like blue
and great tits and ycatchers (Burgess et al., 2018). Phenolog-
ical mismatches are usually magnied up in the trophic chain,
and if few examples have found such effects on large emblem-
atic predators like birds, it is likely that these mismatches also
affect less conspicuous organisms like predatory insects.
Phenological mismatches can also be transmitted by omnivo-
rous animals that can feed on both plants and herbivores, with
potential consequences for the decline of herbivore populations
(Fig. 4d1). The population dynamics of many omnivorous mam-
mals is strongly determined by mast events that often occur peri-
odically (Yang et al., 2010). Oak masting events in the US, for
example, have been found to indirectly dominate the dynamics
of the outbreaking moth Lymantria dispar, which is fed upon by
white-footed mouse, Peromyscus leucopus (Ranesque, 1818),
that feed on oak acorns (Elkinton et al., 1996). Global warm-
ing has been demonstrated to alter the periodicity of oak mating
events (Shibata et al., 2020), which can ultimately alter how gen-
eralist predators regulate herbivore populations. In the L. dispar
example, a reduction in oak masting events could reduce the size
of the populations of omnivorous mammals thus triggering moth
outbreaks, with signicant consequences for other, less compet-
itive insects, that feed on oaks too (Redman & Scriber, 2000).
Similar effects can be expected if global warming alters the peri-
odic appearance of other types of resources like cicadas, whose
adults emerge in large numbers and may alter soil nutritional
composition that cascades up to herbivores via plant growth
(Yang, 2004).
Many populations respond to climate change by expanding
their range with shifts often discordant among species (Gilman
et al., 2010). These shifts in the geographic range of interact-
ing species may drive spatial or temporal mismatches among
these species dramatically altering their interactions (Traill
et al., 2010). Species that expand usually experience a number
of novel abiotic factors in the new ranges. One such factor with
poleward range expansion is that of changes in day length. Day
length can drive interaction strength between species with longer
daylengths altering competitive ability (Kehoe et al., 2018)
and increasing parasitism rate causing co-extinctions through
resource overexploitation (Kehoe et al., 2020).
We are increasingly aware of the importance of insect micro-
bial symbionts for the biology of their hosts (Frago et al., 2020).
Global warming could trigger insect co-extinctions if obliga-
tory symbionts (required for their host survival) are more sen-
sitive to increase temperatures than their hosts, an effect that has
been reported in aphids, stink bugs, whiteies, mealybugs and
weevils (Renoz et al., 2019). Some symbionts, however, while
not required for the survival of their host been found to protect
their hosts against natural enemies, for example, the bacterium
Hamiltonella defensa (Moran, 2005), which protects aphids
from parasitic wasps (Oliver et al., 2009). Symbiont protection
can be lost at elevated temperatures (Doremus et al., 2018), so
that under global warming conditions aphid populations that
rely on this type of protection may be jeopardized, triggering
top-down extinction cascades (Fig. 4d2).
Overexploitation
Although anthropogenic overexploitation is primarily seen
as a problem for megafauna, it being estimated, for example,
to be causing the decline of at least one-third of threatened
birds and amphibians (Navjot et al., 2009), it is also impact-
ing insect species. Overexploitation can directly cause extinc-
tions, as in the case of the recently rediscovered saproxylic
beetle Sclerostomulus nitidus (Benesh, 1955), which exists on
only one mountain. Due to their collection and trade, this bee-
tle has decreased by 93% over a 5-year period (Crespin &
Barahona-Segovia, 2020). The exploitation of edible insects
for large commercial value (as high as $200.00 USD/kg for
the ant Liometopum apiculatum, Mayr, 1870) has decimated
many species (Ramos-Elorduy, 2006). Removing single targeted
species from a community can have far reaching impacts on
communities as shown in Sanders et al. (2015, 2018a). The
harvesting of a single parasitoid species leads to the extinction
© 2020 The Royal Entomological Society, Ecological Entomology, doi: 10.1111/een.12985
10 Rachel Kehoe et al.
of other indirectly linked parasitoids in the experimental insect
communities. The indirect impact depends on the function of the
exploited species, if it has a central role for the community struc-
ture, we can expect a large impact, such as keystone species or
important ecosystem engineers (e.g. ants in the above example).
Conclusions and future directions
Secondary extinctions are likely to play an essential role in
the current decline in insect biodiversity. Most direct impacts
through the anthropogenic drivers have knock-on effects
through co-extinctions and cascading extinctions transmitted
through networks of interactions. So far there is plenty of
evidence for co-extinctions or co-declines, with less research
done on detecting community wide extinction cascades.
Recent theoretical and empirical advances in network ecology
will increase our ability to predict species extinctions. Particu-
larly developments on describing multilayer networks (Pilosof
et al., 2017) and in using molecular biology to identify network
nodes (Hrˇ
cek & Godfray, 2015) are rening the way interac-
tion networks can be constructed. There is also an increasing
interest in studying changes of interaction networks along envi-
ronmental gradients (Pellissier et al., 2018). These studies are
key to understand the potential mechanisms behind extinctions.
For example, network changes along elevational and latitudinal
gradients are useful to assess how networks change with abi-
otic conditions, and to predict extinctions under global warming
scenarios. We are, however, far from similar knowledge with
regards to network changes along gradients of habitat loss (even
if some exist with regards to land use change, see Tylianakis
et al., 2007), dominance of invasive species or intensity of pollu-
tion. These studies are urgently needed to unveil potential indi-
rect mechanisms behind current insect extinctions, but experi-
mental studies are also needed to explicitly test such predicted
mechanisms. To understand how initial impacts are transmit-
ted through the network of species interactions, we need high
quality data on the dynamics of individual species alongside
the structure of the network. These data can then inform theo-
retical models to understand the mechanisms behind extinction
cascades.
The last decade has revealed the importance of microbes
in the biology of animals and plants (Cordovez et al., 2019;
Moran et al., 2019; Frago et al., 2020). As we discussed
in one of the sections, global warming is likely to trigger
insect-symbiont co-extinctions, but we still know little about
how habitat degradation and toxic chemicals, for instance are
altering soil microbiomes with bottom-up effects on plants and
the insects that feed on them. These impacts are also likely to be
strong on freshwater insects, where pollutants may impact them
indirectly via alterations of microbial networks.
Biodiversity loss imposed to natural communities is likely to
be unrepairable in some areas, but the next decade should aim
at reducing extinctions in those areas where natural interaction
networks are still well preserved, and to restore habitats with
dramatic biodiversity declines. To achieve this, it is important
to extend our view from simple pairwise insect extinctions to
those cascading extinctions that can ripple through whole com-
munities. Ecological research on this topic would benet from
more experimental studies, particularly in highly diverse tropi-
cal areas, that are currently under-represented. These will help
us to understand the forces that drive the magnitude of extinc-
tions cascades. Once we develop a predictive understanding, we
can nd strategies to counteract the detrimental impact of initial
biodiversity loss, by either slowing down or stopping cascading
effects to allow the recovery of disturbed ecosystems.
Acknowledgements
We are grateful to Joseph Elkinton and Laura Allen for com-
ments and discussions on the manuscript. EF is currently funded
by the Agence Nationale de la Recherche (ANR) via the PRIMA
S2 2018 project INTOMED and by CIRAD. The authors declare
no conict of interest.
Data availability statement
No data have been used in this research.
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