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Biological Control 75 (2014) 8–17
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Biological Control
journal homepage: www.elsevier.com/locate/ybcon
Relationships between biodiversity and biological control in
agroecosystems: Current status and future challenges
David W. Crowder a,⇑, Randa Jabbour b
a
b
Department of Entomology, Washington State University, PO Box 646382, Pullman, WA, USA
Department of Plant Sciences, University of Wyoming, 1000 E. University Ave., Laramie, WY 82071, USA
h i g h l i g h t s
g r a p h i c a l a b s t r a c t
Agricultural systems are intensifying
to keep up with growing human
demands.
Intensive agriculture can negatively
impact biodiversity and biological
control.
We review biodiversity and biological
control of arthropod and weedy pests.
We discuss similarities and
differences between biocontrol of
arthropods and weeds.
We suggest novel approaches for
examining biodiversity and biological
control.
a r t i c l e
i n f o
Article history:
Available online 1 November 2013
Keywords:
Agricultural intensification
Conservation
Ecosystem functioning
Evenness
Richness
a b s t r a c t
Agricultural systems around the world are faced with the challenge of providing for the demands of a
growing human population. To meet this demand, agricultural systems have intensified to produce more
crops per unit area at the expense of greater inputs. Agricultural intensification, while yielding more
crops, generally has detrimental impacts on biodiversity. However, intensified agricultural systems often
have fewer pests than more ‘‘environmentally-friendly’’ systems, which is believed to be primarily due to
extensive pesticide use on intensive farms. In turn, to be competitive, less-intensive agricultural systems
must rely on biological control of pests. Biological pest control is a complex ecosystem service that is generally positively associated with biodiversity of natural enemy guilds. Yet, we still have a limited understanding of the relationships between biodiversity and biological control in agroecosystems, and the
mechanisms underlying these relationships. Here, we review the effects of agricultural intensification
on the diversity of natural enemy communities attacking arthropod pests and weeds. We next discuss
how biodiversity of these communities impacts pest control, and the mechanisms underlying these
effects. We focus in particular on novel conceptual issues such as relationships between richness, evenness, abundance, and pest control. Moreover, we discuss novel experimental approaches that can be used
to explore the relationships between biodiversity and biological control in agroecosystems. In particular,
we highlight new experimental frontiers regarding evenness, realistic manipulations of biodiversity, and
functional and genetic diversity. Management shifts that aim to conserve diversity while suppressing
both insect and weed pests will help growers to face future challenges. Moreover, a greater understanding of the interactions between diversity components, and the mechanisms underlying biodiversity
effects, would improve efforts to strengthen biological control in agroecosystems.
Ó 2013 Elsevier Inc. All rights reserved.
1. Introduction
⇑ Corresponding author.
E-mail address: [email protected] (D.W. Crowder).
1049-9644/$ - see front matter Ó 2013 Elsevier Inc. All rights reserved.
http://dx.doi.org/10.1016/j.biocontrol.2013.10.010
Human population growth has led to the global expansion of
agriculture. The acreage of land used for crops increased by 466%
D.W. Crowder, R. Jabbour / Biological Control 75 (2014) 8–17
from 1700 to the 1980s (Meyer and Turner, 1992). This growth,
however, has slowed in recent decades as suitable areas for cultivation become increasingly scarce (Matson et al., 1997). As growth
of agricultural acreage has stagnated, agricultural systems around
the world have intensified. Agricultural intensification is a broad
term that encompasses many factors including, but not limited
to, increased use of pesticides and fertilizers (see Roubos et al.,
2014), increases in farm size, decreases in crop diversity, increases
in crop density, and increased numbers of crops grown in a season.
This has been due in large part to dramatic increases in crop yields
since the 1960s, referred to as the ‘‘Green Revolution’’ (Pingali,
2012), which has been spurred by technological and cultural advances in crop breeding and management (Matson et al., 1997;
Krebs et al., 1999; Benton et al., 2003).
While agriculture has kept pace with human population
growth, increases in crop yields has also slowed recently (FAO,
2013). Moreover, agricultural intensification has negative local
consequences such as reduced biodiversity, increased soil erosion,
pollution, and reduced socio-economic sustainability, each of
which has other impacts (Matson et al., 1997; Stoate et al., 2001;
Kleijn et al., 2006). For example, reducing the number of species
(reduced richness) (Hooper et al., 2005; Cardinale et al., 2006)
and skewed relative abundance distributions (reduced species
evenness) (Hillebrand et al., 2008; Crowder et al., 2010) generally
weaken biological control. These harmful consequences of agricultural intensification have led to an increased focus on methods to
increase the sustainability of agroecosystems (Tilman, 1999; Foley
et al., 2011).
Biological control is a key ecosystem service that is necessary
for sustainable crop production (Bianchi et al., 2006; Losey and
Vaughan, 2006). Natural enemies such as predators, parasitoids,
and pathogens play a central role in limiting damage from native
and exotic pests (Hawkins et al., 1999; Losey and Vaughan,
2006). Conservative estimates suggest that the economic value
provided by insect natural enemies controlling pests attacking crop
plants exceeds $4.5 billion annually in the United States (Losey and
Vaughan, 2006). If weedy pests, or pests attacking humans and
livestock (not crops) were considered the estimate of pest control
provided by insects would likely be much greater. Moreover, a
multitude of species act as natural enemies of insect or weedy
pests such as birds, bats, fungi, nematodes, and rodents (Kirk
et al., 1996; Miller and Surlykke, 2001; Navntoft et al., 2009; Ramirez and Snyder, 2009; Williams et al., 2009; Jabbour et al.,
2011). Thus, if these species were considered the economic value
of biological control would be far greater than $4.5 billion annually.
To improve and conserve biological control, it is essential to
understand the relationships between agricultural intensification,
biodiversity, and pest suppression. We addresses this complex issue by first reviewing the effects of agricultural intensification on
the biodiversity of natural enemies, and the role of natural enemies
in agricultural food webs. We next discuss conceptual models
relating biodiversity to natural pest control. Third, we review
methodologies for examining the relationship between biodiversity and biological control in agroecosystems. We conclude by discussing areas of research emphasis that, if addressed, would
improve our understanding of how biodiversity and biological control operate in agroecosystems.
2. Effects of agricultural intensification on biodiversity
2.1. Non-pest species
Agricultural intensification impacts both pest and non-pest species in agricultural communities. Indeed, much of the evidence of
the impacts of agricultural intensification on ecological
communities comes from conservation-related studies on non-pest
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species. Some of the longest-term studies of agricultural intensification and biodiversity have focused on bird populations in Europe,
which have declined dramatically over the last half-century
(Benton et al., 2003). Donald et al. (2001) showed that bird
populations in the UK declined with increases in cereal and milk
yields along with fertilizer and tractor usage. Cereal yields alone
explained 31% of the variability in declining bird populations,
suggesting that intensification of a single crop type can impact
diversity (Donald et al., 2001).
There is evidence, however, that agri-environment schemes enacted by many European countries to encourage wildlife have led
to resurgence of some bird species (Benton et al., 2003). Agri-environment schemes provide monetary incentives for farmers to manage a portion of their land to promote conservation of biodiversity
and reduced impacts on the environment (Benton et al., 2003;
Kleijn et al., 2006). By incorporating or conserving natural habitat
in agricultural ecosystems to preserve native species, these
schemes are designed to buffer against potentially damaging
effects from agricultural intensification on biodiversity. Kleijn
et al., 2006 compared the abundance and richness of plants, birds,
and arthropods at 202 paired locations across five European
counties. Each location contained one site managed with an
agri-environment scheme and one conventional site. The agrienvironment schemes had some positive impacts on abundance
and diversity of these groups in each country, while conventional
management did not benefit any group (Kleijn et al., 2006). The
authors speculated that benefits were due to reduced inputs and
disturbances in agri-environment fields. However, the species that
benefitted most from agri-environment schemes did not include
many species of extinction concern. This suggests that conserving
native habitat may not benefit rare species, or that species of
extinction concern declined in abundance due to factors other than
agricultural intensification.
Crowder et al. (2010, 2012) showed that organic farming systems had marginal positive impacts on richness and significant positive impacts on evenness and abundance compared with
conventional systems. These benefits occurred across crop types
and were consistent across several organismal groups including
arthropods, birds, non-bird vertebrates, plants, and soil organisms.
The benefits of organic farming were greatest for the rarest species
in conventional systems (Crowder et al., 2012). Other reviews
(Bengtsson et al., 2005; Hole et al., 2005) have shown similar positive impacts of organic farming on richness and abundance of
organisms. In each case, these results are likely due to reduced
insecticide use in organic farming systems and/or increased habitat
diversity. For example, granivorous beetle diversity has been
shown to be positively associated with habitat complexity on
farms (Vanbergen et al., 2010; Trichard et al., 2013), and negatively
associated with use of pesticides (Trichard et al., 2013).
2.2. Arthropod pests
Root (1973) suggested that dense, homogenous plant communities facilitated higher herbivore populations. His ‘‘resource-concentration hypothesis’’ posits that specialist pests can locate
plant stands, and feed more efficiently, when a single non-diverse
crop is present. Thus, intensification may actually be responsible
for pest outbreaks so common in monocultures. However, this
hypothesis does not hold true for all cases, suggesting that resource-concentration effects are organism dependent (Grez and
Gonzalez, 1995). Secondary pest outbreaks, where early-season
insecticide applications kill natural enemies and result in late-season outbreaks of pests, have also received attention as a negative
impact of agricultural intensification. For example, Gross and
Rosenheim (2011) showed that 20% of late-season pesticide costs
were attributable to early-season pesticide applications for control
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of lygus bugs. This is perhaps not surprising as pesticides can disrupt natural enemy populations at broad scales (Roubos et al.,
2014) and might disrupt periods of temporal overlap between
pests and their natural enemies (Welch and Harwood, 2014).
Comparison of organic and conventional farming systems also
suggest mixed effects of intensification on pests. For example,
aphids tend to be more abundant in conventional farms due to
higher-quality crops that grow faster, while mites and true bugs
tend to be more common in organic farms (Hole et al., 2005).
Across multiple pest groups, Bengtsson et al. (2005) found no significant effect of organic farming on abundance or species richness.
This result is perhaps surprising given that conventional farms
tend to use significantly more pesticides compared with organic
farms, which should limit pest populations (Bengtsson et al.,
2005; Hole et al., 2005). This perhaps suggests that for many pest
herbivores that increased effectiveness of natural enemies in organic farming systems might allow for pest control equivalent, or
nearly equivalent, to control achieved with pesticides (Crowder
et al., 2010, 2012).
2.3. Weeds
Weed communities are consistently more abundant and rich in
less-intensive agricultural systems. For example, organic farms
generally harbor 2.3–2.8 times more abundant weed seed densities
and 1.3–1.6 times more weed species compared with conventional
farms (Roschewitz et al., 2005; Gabriel et al., 2006; Hawes et al.,
2010; José-María and Sans, 2011). Frequent herbicide use in intensive agroecosystems likely drives reduced weed diversity and
abundance. In addition, weed communities often vary based on
inorganic fertilizer use and crop rotations (Hawes et al., 2010).
Hawes et al. (2010) showed that fertilizer use and rotations explained as much variation in weed abundance/ diversity as farm
type across 109 conventional, integrated, and organic farms. Synthetic fertilizer use was negatively associated with weed abundance, an unsurprising result given that farms using synthetic
fertilizers also likely use synthetic herbicides. However, organic
manures are known as likely sources of weed seed (Mt. Pleasant
and Schlather, 1994), a risk that farmers and scientists acknowledge readily in discussions of weed introduction and weed spread
(Jabbour et al., 2013).
Weed communities on farms are also influenced by the surrounding landscape (see also Chisholm et al., 2014). For example,
weed vegetation and seedbank diversity in German wheat fields
was positively associated with landscape complexity around farms
(Roschewitz et al., 2005). Interestingly, in this study, weed diversity was similar in organic and conventional fields when landscapes were complex, but weed diversity was higher in organic
fields in simple landscapes. In contrast, effects of landscape complexity on weed seedbanks in Mediterranean dryland systems
were limited and only detected in field edges, not field centers
(José-María and Sans, 2011). These results suggest that in some,
but not all systems, management of landscape complexity might
be a strategy for less-intensive farmers to manage weeds.
2.5. Overview
The studies discussed here show that agricultural intensification has complex effects on pests and beneficial species. In general,
more intensive systems decrease the abundance and biodiversity
of beneficial species such as natural enemies. At the same time,
intensification might exacerbate pest problems by concentrating
arthropod resources and decreasing populations of natural enemies. In the following sections we discuss how disruptions in biodiversity of pests and natural enemies caused by agricultural
intensification can influence species interactions and have strong
impacts on biological control.
3. Why does biodiversity matter? Exploring agricultural food
webs
Reductions in biodiversity can disrupt food webs and weaken
biological control (Tylianakis et al., 2007; Macfadyen et al., 2009;
Schmitz and Barton, 2014; Tylianakis and Binzer, 2014). Trophic
cascades have been a central concept in biocontrol studies since
Hairston et al. (1960) proposed the ‘‘Green World Hypothesis’’,
whereby natural enemies protect plants by regulating pests. However, the role of biodiversity in generating trophic cascades has
been questioned (Strong, 1992; Polis and Strong, 1996; Polis,
1999). A key criticism is that in complex food webs, removal of a
single species is unlikely to influence cascades. Indeed, studies
have shown that natural enemy diversity can promote (Cardinale
et al., 2003; Snyder et al., 2006; Crowder et al., 2010) or weaken
trophic cascades (Finke and Denno, 2004, 2005). Understanding
the food webs in which natural enemies and pests interact is thus
central to understanding relationships between biodiversity and
biological control.
Arthropod food webs are discussed by Tylianakis and Binzer,
2014, and we do not expound on them here. Weeds also play key
roles in agricultural food webs and may thus influence biodiversity
and biological control (Fig. 1). Farmers generally aim to manage
weeds given their considerable agricultural and economic risks,
although they do acknowledge the ecological benefits of weeds
(Wilson et al., 2008; Jabbour et al., 2013). As hypothesized by Root
(1973), plant diversity is often negatively associated with herbivore abundance (Cardinale et al., 2012). Weeds comprise a large
component of floral diversity in agroecosystems, given the comparably low diversity of most crop plantings, and thus may play a key
role in limiting pests from the bottom-up.
2.4. Other pests
There is evidence that agricultural intensification influences
pathogens (Wolfe et al., 2007). For example, measles, mumps,
smallpox, and influenza likely originated with domesticated animals (Wolfe et al., 2007). Agriculture leads to crowding of populations, which promotes pathogen spread (Wolfe et al., 2007).
Similarly, high vector densities in intensive systems promote the
spread of crop pathogens. While not a focus of this review due to
limited studies, effects of biodiversity on biological control of
pathogens should be examined further.
Fig. 1. A hypothetical agricultural food web, where arrows indicate energy flow.
Biological control of weeds and insect prey involve several of the same animal
groups, including potential intraguild predation amongst birds, mammals, and
insect omnivores and predators.
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In addition to the potential contribution of weed diversity to
limiting pests, the resource of weed seeds in particular can have
major impacts on food webs. Weed seed resources can support a
high diversity of granivorous and omnivorous species including
mammals, birds, arthropods, earthworms, and slugs (Franke
et al., 2009). A detailed study of the seed-based food web on an organic farm in the UK estimated that 274 arthropod, 53 bird, and 10
mammal species likely used seeds as a resource (Evans et al.,
2011). However, these effects can be quite nuanced and dependent
on crop type. A recent study focused on farmland birds showed
that genetically-modified beet and oilseed crop systems had less
weed seed resources used by birds than conventional systems
(Gibbons et al., 2006). In contrast, genetically-modified maize
had more seed available for birds than conventional systems.
Although seed resources can support omnivores, seed availability
may also distract from insect pest control by natural enemies
(e.g., Frank et al., 2011).
These studies demonstrate that natural enemy consumers and
their resources (arthropod and weedy pests) interact in complex
food webs (see also Tylianakis and Binzer, 2014). Food-web ecology is a burgeoning field that could thus contribute to our understanding of the relationship between biodiversity and biological
control. We expound on this issue in the following section.
4. Relating richness to biological control
4.1. Mechanisms underlying effects of richness on biological control
The mechanisms by which species richness affects biological
control of arthropods have been reviewed (Schmitz, 2007; Letourneau et al., 2009), and we only briefly discuss them to provide context for the remainder of the review (Fig. 2). Positive richness
effects occur when species act complementarily in terms of pest
suppression, or when one or more species facilitates prey capture
by another, such that the combined impact of multiple species exceeds the mortality any species exerts on its own (Losey and Denno, 1998; Finke and Snyder, 2008; Northfield et al., 2010). Positive
richness effects can also occur through ‘‘insurance effects’’, where a
more species rich community effectively performs biological control despite disturbances because one or more species is resilient
to a disturbance even while others are negatively affected (Naeem
and Li, 1997; Yachi and Loreau, 1999). In contrast, increasing species richness can negatively impact biological control when natural
enemies feed on each other, an occurrence known as intraguild
predation, which limits predator density and impact on pests
(Finke and Denno, 2004, 2005; Rosenheim, 2007). Behavioral interference among predator species can also weaken biological control
with increases in natural enemy richness (Schmitz, 2007).
Many authors have also pointed out that the positive or negative impacts of species richness on ecosystem functions such as
biological control can be due to statistical chance. In this scenario,
richer communities could improve biological control simply because they have a higher probability of containing a superior natural enemy species (Naeem and Wright, 2003; Cardinale et al.,
2006). In contrast, richer communities might also have a higher
probability of containing a disruptive species (like a voracious
intraguild predator), which could weaken biological control
(known as a selection effect; Loreau and Hector, 2001).
4.2. The richness-biological control relationship in studies involving
arthropod pests
Studies relating richness to arthropod biological control have
variably shown positive, negative, and neutral effects. In a metaanalysis, Letourneau et al. (2009) showed that 71% of studies
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involving arthropod pests observed positive effects of natural enemy richness on pest suppression, with particularly strong effects
in agricultural systems. Nearly all of the remaining 30% of studies,
however, showed a negative relationship between natural enemy
richness on pest suppression. Yet, the magnitude of the average
negative effect was 63% that of the average positive effect, suggesting that natural enemy richness was more likely to generate strong
positive impacts than strong negative effects. Letourneau et al.
(2009) could not distinguish between the mechanisms underlying
these effects of richness, however. Thus, it was unclear whether
more diverse communities promoted pest control because natural
enemies tended to act complementary or facilitated prey-capture
by other species, or whether more diverse communities were more
likely to contain particularly effective natural enemy species by
chance alone (otherwise known as a sampling or species identity
effect).
4.3. The richness-biological control relationship in studies involving
weedy pests
Evidence for richness effects on biological control of weeds is
less common than for arthropods. However, in Wisconsin and
France, richer seed-feeding beetle communities were associated
with increased weed seed loss from experiments (Gaines and Gratton, 2010; Trichard et al., 2013). This positive relationship occurred
in both crop and non-crop habitats in Wisconsin potato agroecosystems (Gaines and Gratton, 2010). As with arthropods, indirect
evidence suggests that complementarity may underlie these effects. Invertebrate granivores eat smaller seeds on average compared to mammals and birds, such that the combined impacts of
both guilds of natural enemies strengthened weed suppression in
the Gaines and Gratton study (2010). Different birds species vary
in their weed-seed preferences as well (Perkins et al., 2007). Weed
predators also may be temporally complementary, with invertebrate seed consumption decreasing and vertebrate consumption
increasing as winter approaches (Davis and Raghu, 2010).
Alternatively, seed consumer richness may negatively affect
weed seed predation via intraguild predation. Many granivores in
the seed-based food web are omnivores who also prey upon invertebrates. Small mammals feed on insects as well as seeds, and
could negatively impact invertebrate granivores. For example,
excluding rodents from experimental plots in shrub-steppe habitats resulted in increased carabid abundance following a two year
period (Parmenter and MacMahon, 1988). Moreover, a study of
invertebrate and vertebrate predation in corn fields highlighted
the potential influence of predation risk of seed predators on seed
predation rates (Davis and Raghu, 2010). Invertebrate seed predation was negatively associated with abundance of spiders, potential predators of granivorous crickets active in that system. Also,
vertebrate seed predation rates were positively associated with
cloud cover; vertebrates may avoid foraging in cropping fields on
clear nights due to increased predation risk (Davis and Raghu,
2010).
4.4. Overview
Studies involving arthropods and weeds have shown variable
effects of richness on biological control. However, most studies
have not examined other aspects of biodiversity such as evenness,
genetic diversity, or functional trait diversity, and mechanisms are
largely inferred rather than demonstrated. Research is needed that
expands beyond richness effects and develops sophisticated approaches for isolating mechanisms underlying biodiversity effects.
We discuss these two issues in the remainder of this review.
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Fig. 2. Effects of richness on biological control, where the amount of leaf surface covered by predators correlates with the number of prey captured. In (A,B), the predators
forage in different locations, and more prey are consumed across the whole leaf when both predators are present (A) compared to only one predator (B). In (C–F), the
predators overlap in their distribution. If predators are redundant (C–E) the same amount of prey are captured regardless of whether one (D) or both (C,E) predators are
present. However, if intraguild predation occurs (F), total predator density decreases and biological control weakens (F).
5. Novel research concepts: moving beyond studies of richness
and biological control
Most studies involving biodiversity and biological control of
arthropod and weed pests have examined the effects of species
richness. Indeed, the terms ‘‘biodiversity’’ and ‘‘richness’’ are often
used synonymously by many authors, and incorrectly when
authors equate richness to total diversity of a system. In the following section we describe new frontiers in biodiversity and biological control research that should be addressed to move
beyond an understanding of richness effects and increase the relevance for real-world agroecosystems.
5.1. Effects of evenness
Evidence is accumulating that evenness can have similar effects
to richness on a variety of ecosystem processes including biological
control (Hillebrand et al., 2008). Crowder et al. (2010) showed that
relatively small increases in evenness of predator and pathogen
communities significantly improved control of the Colorado potato
beetle Leptinotarsa decemlineata (Coleoptera: Chrysomelidae), generating a strong trophic cascade and larger potato plants. The
authors showed that intraspecific competition was reduced in
more even communities, leading to increased natural enemy survival in even communities and greater pest control.
Fig. 3 shows a conceptual model for how evenness might influence pest control. When natural enemy species are complementary, a more even community captures more prey (A). This
occurs because when the community is uneven, the rare species
does not saturate its’ resource niche, such that total prey consumed
decreases (B). In contrast, when natural enemies are redundant,
evenness does not affect pest control because any combination of
individuals captures the same number of prey (C,D). When predators differ in their effectiveness at biological control, evenness
could have negative or positive impacts. In this case, when the
community is uneven and the more voracious predator is dominant, prey consumption increases compared with the even community (E,F). However, if a community is uneven and the more
voracious predator is rare, evenness would have a positive effect
on pest control.
This conceptual model represents some of the possible effects of
evenness on pest control, and is not meant to be exhaustive. In
each scenario, if predator density was infinitely high, evenness
would not matter, as even the rare species would saturate its’
niche. However, species rarely exist at high-enough densities to kill
all prey available to them (Northfield et al., 2010), such that effects
of evenness on prey consumption in most conditions will be mediated by abundance. To our knowledge no study has varied density
and evenness in a systematic manner to test these predictions. Future research in this area is therefore warranted to uncover scenarios where evenness will have positive, negative, or neutral impacts
on biocontrol.
5.2. Realistic manipulations of biodiversity
One drawback of experiments relating biodiversity to ecosystem functions has been a lack of realism (Bracken et al., 2008;
Byrnes and Stachowicz, 2009). In most experiments, treatments
consist of single species in monoculture or diverse communities
with even distributions, such that realism is sacrificed to increase
rigor and interpretability. Communities in agroecosystems, however, do not vary in such predictable manners (Crowder et al.,
2012). Moreover, the effects of biodiversity on ecosystem functions
can differ significantly in randomly assembled communities compared with non-random assemblages reflecting real-world variation (Bracken et al., 2008). As described earlier, Crowder et al.
(2010) overcame this problem by establishing evenness treatments
that reflected variation in species evenness on farms. In this way,
biological control and plant densities in experimental trials were
informative for potential impacts on farms. While these
D.W. Crowder, R. Jabbour / Biological Control 75 (2014) 8–17
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Fig. 3. Effects of richness on biological control, where the amount of leaf surface covered by predators correlates with the number of prey captured. In (A,B), the predators
forage in different locations, and more prey are captured when the community is even (A) than when it is uneven (B). In the latter scenario, there are too-few predators
feeding on the leaf edge to effectively capture prey in this spatial niche. In (C,D), the predators overlap in their distribution, and are redundant, such that evenness does not
impact biological control. In (E,F) the predators overlap in their distribution, but one is more effective at prey capture than the other. In this case, the uneven (F) community
can outperform the even (E) community when the more effective predator is dominant, although the uneven community will underperform if the more effective predator is
not dominant (not shown).
approaches add realism to experiments testing biological control, a
potential drawback is that realistic biodiversity designs can be statistically messy due to covariance in species abundances in managed ecosystems. Crowder et al. (2010) dealt with this issue by
including in statistical models the abundance of each predator species in addition to biodiversity, and suggested that information
theoretic criterion could be used to distinguish between effects
of particular species and effects of biodiversity per se. Despite these
challenges, more experiments should attempt to realistically vary
biodiversity to improve our understanding of how biological control operates in real-world agroecosystems.
Diversity at multiple trophic levels also occurs in agroecosystems. For example, rarely do agricultural systems contain a single
pest or plant species, as weeds and crop plants are often intermixed. This variation can affect the biodiversity-biological control
relationship. For example, intraguild predation amongst generalist
predators could be lessened by the availability of multiple prey
species, although this could also weaken biological control of a focal species (Frank et al., 2010, 2011). Frank et al. (2011) showed
that seed subsidies in open-field plots increased the abundance
of omnivorous carabids in corn, but there was also more cutworm
damage documented. In general, however, few studies have examined effects of diversity at multiple trophic levels on biological control, even though predators, herbivores, and plants co-exist in
intricate food webs (Tylianakis and Binzer, 2014). Future research
in this area could improve our understanding of realistic variation
throughout food webs on biological control.
5.3. Functional diversity
Functional diversity, rather than biodiversity per se, may often
underlie the biodiversity-biological control relationship. When
species are lumped into functional groups based on ecological
traits, species in any group are considered ecologically redundant
and species in different groups are complementary (Hillebrand
and Matthiessen, 2009). Species can be functionally grouped based
on resource/habitat preferences and seasonal differences (Northfield et al., 2012). Increased numbers of functional groups are expected to increase biological control, while increased diversity
within a functional group would not (Northfield et al., 2012). For
example, on ragwort plants, moth larvae feed on leaves while flea
beetles tunnel into leaf petioles and roots (McEvoy et al., 1991).
These two herbivores therefore represent two functional feeding
groups, and their combined presence improves ragwort control
(James et al., 1992). A meta-analysis to test whether multiple natural enemies used to control weeds had independent or non-independent effects on plant performance showed that, indeed,
antagonistic effects were likely to occur if both enemies were
attacking the same plant part (Stephens et al., 2013). However,
grouping species into functional groups can be difficult because
we often do not have sufficient ecological information to create
functional groups of broad relevance for ecosystem services such
as biological control (Wright et al., 2006).
Understanding functional diversity would improve our ability
to conserve natural enemy communities to improve biological control. If managers could predict which functional groups are important, they could conserve functional diversity rather than diversity
per se. Weed diversity can also be partitioned into functional
groups that are more relevant to farmers. For example, weed species vary throughout the season in timing of germination and seed
rain (e.g., cool season vs. warm season weeds), factors that relate to
timing of herbicide, cultivation, planting, and harvest decisions.
Functional diversity can also be important when considering temporal dynamics of arthropod pests, as natural enemy species with
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D.W. Crowder, R. Jabbour / Biological Control 75 (2014) 8–17
different temporal dynamics (i.e., species that attack in different
parts of a day or in a season) may be necessary to achieve adequate
pest control (Welch and Harwood, 2014). Annual and perennial
weeds require distinct management approaches, given the ability
for some perennials to propagate vegetatively, particularly problematic for organic farmers (e.g., Turner et al., 2007).
However, the utility of functional diversity schemes has been
questioned, because classifications often fail to explain ecological
processes (Wright et al., 2006). For example, assigning plants to
random functional groups explained as much variation in biomass
and nitrogen assimilation as when plants were placed in well-defined groups (grasses, forbs, legumes and woody plants). Schmitz
(2007) provided one of the most detailed descriptions of functional
diversity in a system of spiders feeding on grasshopper prey. In this
scheme, Schmitz defined functional groups based on habitat domains, which encompass the microhabitats chosen by natural enemies and the extent of spatial movement in those microhabitats,
and showed these groupings had strong predictive ability for predator–prey relationships (Schmitz, 2008). While informative, such
schemes require detailed ecological information, and would likely
need to be developed on a case-by-case basis. Moreover, several
studies have suggested that climate change might shift the functional roles of species in ecosystems, changing functional groupings and food-web interactions (Barton and Schmitz, 2009;
Schmitz and Barton, 2014; Tylianakis and Binzer, 2014) in addition
to altering temporal interactions between species (Welch and Harwood, 2014). This will complicate our understanding of how biological control might operate during periods of climate change.
5.4. Genetic diversity
Genotypic diversity can have strong consequences for ecological processes (Hughes et al., 2008; Cook-Patton et al., 2011). However, few studies have examined effects of genetic diversity on
biological control. Studies from plants indicate that variation in genetic diversity within plant species can have similar effects compared with species diversity on ecosystem functions such as
biomass production (Schweitzer et al., 2005; Crutsinger et al.,
2006; Johnson et al., 2006; Cook-Patton et al., 2011). Complementarity has also been shown to drive positive effects of genetic diversity, similar to species-diversity studies (Cook-Patton et al., 2011).
There is evidence that genetic diversity within species that act as
biological control agents results in variation in climatic tolerances,
prey and habitat preferences, and synchrony with hosts among
others (Hopper et al., 1993). Hopper et al. (1993) showed up to
14-fold variation in these traits within populations of a single natural enemy species. This suggests that genetic diversity within natural enemy species, and impacts on biological control, should be a
priority area for future research.
5.5. Overview
Considerable attention has been paid to the role of natural enemy species richness in biological control. However, the relevance
of studies examining biodiversity, biological control, and realworld management could be improved by further examining other
dimensions of biodiversity in agroecosystems. In particular, we
highlight the potential importance of evenness, functional diversity, and genetic diversity. These facets of biodiversity consistently
vary across real-world agroecosystems, and uncovering the mechanisms through which these factors impact biological control is a
significant challenge for ecology. One step in the right direction
might be moving towards experiments that use realistic manipulations of biodiversity to improve our understanding of processes in
real-world ecosystems.
6. Elucidating mechanisms underlying the biodiversitybiological control relationship
Ecologists have struggled to identify the reasons that biodiversity might influence ecosystem functions like biological control
(Naeem and Wright, 2003; Cardinale et al., 2006). For example, positive effects could be driven by complementary (where species attack prey in unique ways) or facilitation (where one natural enemy
might increase prey-capture by another species). Negative effects
of biodiversity might be driven by factors such as intraguild predation, where different natural enemies feed on each other rather
than pests. One roadblock has been that many traits differ among
species such as size, feeding preferences, and metabolism (Huston,
1997; Loreau and Hector, 2001). Thus, many traits are varied
whenever species diversity is manipulated, making it difficult to
identify how a single trait like complementary niches might influence biological control (Finke and Snyder, 2008; Northfield et al.,
2010). One approach to separate positive or negative species interactions from identity effects is to determine whether ecological
performance of diverse communities exceeds each single species
(Petchey 2003). This approach, however, fails to identify mechanisms underlying effects of biodiversity like complementarity or
intraguild predation (Petchey 2003). Models have also been used
to test for effects of diversity in conjunction with species identity
effects (Crowder et al., 2010), yet this method also fails to identify
mechanisms. Here we discuss novel methods that can directly test
specific mechanisms underlying the relationship between biodiversity and biological control.
6.1. Linking experiments with models
Theoretical models have provided insight into the mechanisms
underlying the effects of biodiversity on biological control (Sih
et al., 1998; Ives et al., 2005; Casula et al., 2006). However, models
of biodiversity and biological control have rarely been confronted
with data. Northfield et al. (2010) showed how models could be
linked with a complex experiment to explore the relationship between biodiversity and biological control. This study relied on a response-surface design (Inouye, 2001; Paini et al., 2008) whereby
total abundance and species diversity were varied systematically.
Response-surface designs allow for isolation of interspecific interactions characteristic of additive designs (Sih et al., 1998; Ives
et al., 2005), while also taking advantage of substitutive design
which eliminate confounding effects of abundance across treatments (Connolly 1988; Hooper et al., 2005).
The Northfield et al. (2010) study varied densities and species
richness of a community of predators attacking aphids on collard
plants. The authors took advantage of the concept of niche saturation, whereby a single predator species is only capable of consuming a proportion of the total prey base. As abundance of a predator
species increases, they consume all of their available prey, and
thereby saturate their niche, where the proportion of prey remaining is the proportion outside of the predator’s niche. Thus, if a second predator can consume additional prey, it indicates they were
able to consume prey outside of the first predator’s niche, indicating a complementary feeding interaction. Using this framework,
Northfield et al. (2010) evaluated their experimental results with
a model that extended upon the framework developed by Casula
et al. (2006). In the model, parameters were included for niche partitioning, facilitation, and niche overlap. The authors used likelihood-ratio tests to selectively explore whether each factor
improved the fit of the model to the data. In this way, Northfield
et al. (2010) were able to convincingly demonstrate that niche partitioning was a key factor driving positive biodiversity effects on
aphid biological control (measured in reduction of aphid numbers).
D.W. Crowder, R. Jabbour / Biological Control 75 (2014) 8–17
The framework developed by Northfield et al. (2010) is general,
and could be applied to other systems. Moreover, the model would
allow for variation in richness and evenness, while still having the
capacity to test mechanisms. Future studies should use this framework and/or expand on it to test mechanisms driving biodiversity
effects on biological control.
6.2. Manipulations of niche-breadth
Several authors have directly demonstrated complementarity
by directly manipulating niche breadth of species independent of
species identity. One of the best examples comes from the study
of Finke and Snyder (2008). In this study, three species of parasitoids were ‘‘trained’’ to use a single species of aphids (i.e., they were
specialists) or multiple aphid species (i.e., they were generalists) as
hosts. The authors then directly manipulated the niche overlap and
species diversity independently within the parasitoids communities. They showed that only when aphids were specialists and used
separate host resources did increased biodiversity improve biological control, directly demonstrating that niche complementarity
improved biological control.
Gable et al. (2012) used a different approach, whereby ecological engineers were used to naturally manipulate plant habitats and
niche relationships among species. In this study, conducted on collard plants, lady beetle predators typically foraged only on leaf
edges while parasitoids and true bug predators foraged in leaf centers. In turn, when leaves were intact, this spatial niche partitioning among species led to an improvement in biological control only
when a diverse community of natural enemies was present. However, when diamondback moths were present on leaves, they
chewed holes in leaves, which allowed lady beetles to forage over
the entire plant surface. In this case, increases in biodiversity did
not improve biological control. Thus, by directly manipulating spatial niche partitioning among predators, the authors showed that
spatial complementarity could explain the biodiversity-biological
control relationship.
An advantage of these approaches is that they might reveal how
diversity effects naturally vary across ecosystems, with species
complementing one another only in certain situations. The drawback is that they require quite a detailed understanding of species
traits, so complementarity can be directly manipulated. Despite
this challenge, more studies in this vein would help uncover mechanisms driving the relationship between biodiversity and biological control.
15
insects, and managing for conservation of diverse groups that contribute to this function such as invertebrates, mammals, and birds.
Managers can work to restore diversity through a variety of
manipulations.
For example, crop rotation has nutrient and pest management
benefits, and serves as a strategy to diversify agricultural systems.
A comparison of rotation systems in Iowa, ranging from 2 to 4 year
sequences with varying levels of synthetic inputs, demonstrated
that weeds were suppressed in all systems, and yield and profit
was the same or improved in diversified rotations as compared
to the conventional 2-year corn-soybean rotation (Davis et al.,
2012). Weeds in the diversified rotations were managed using a
variety of strategies, effective by targeting weeds at different
stages in the life cycle and season (Liebman and Gallandt, 1997).
Furthermore, crop rotations in these areas were effective at controlling corn rootworm pests that lay eggs in corn fields, as eggs
hatching in soybean fields the following year die (Onstad et al.,
2003).
One notable result of agricultural intensification and specialization has been the decoupling of crop and livestock production systems. The availability and low cost of synthetic fertilizer has
enabled crop production without the need for animal manure,
and efficient feedlot technology and inexpensive feed has enabled
livestock production separate from cropland. Re-integrating crop
and livestock systems has been proposed as a solution to many
environmental problems of modern agriculture, such as nitrogen
leaching, poor soil quality, and manure management (Russelle
et al., 2007). Re-integration also has the potential to improve pest
control via biological control in the form of grazing. In Montana,
sheep grazing of dryland grain fallow and alfalfa contributed to
management of wheat steam sawfly and alfalfa weevil, major pests
in these respective systems, and weedy plants (Goosey et al., 2005;
Goosey, 2012). Although this strategy may be logistically and economically challenging for a single grower to accomplish, partnerships amongst growers within a local area could lead to
successful integration.
Crop rotations and coupling crop and livestock production are
just two of many potential options for growers to potentially conserve biodiversity and promote ecosystem services like biological
control. To truly promote sustainability in our agroecosystems, it
will be essential for managers to consider multiple aspects of communities such as richness, evenness, abundance, and species identity to maximize potential benefits for farming systems. While this
is a daunting challenge, it is one we must overcome for agriculture
to continue to meet the demands of a growing human population
in an ever-changing world.
7. Managing for biodiversity and biological control
Acknowledgments
Overall, agricultural managers face an uncertain future of more
extreme and changing climatic conditions, depleting water supply,
and increased costs of fossil fuels. In this review we show that
diversification of agroecosystems generally decreases impacts on
natural resources, promotes more diverse natural enemy communities, and strengthens biological control. While agricultural intensification is likely to continue, it is clear that future farming
systems will have to consider biodiversity and natural ecosystem
functions such as biological control to continue to meet the global
challenges facing agriculture in a sustainable and efficient manner.
One central management challenge that this paper poses is
‘‘how can we manage for diversity while suppressing both insect
and weedy pests?’’ Insects and weeds are rarely considered in concert by researchers, although growers of course face both every
day. As demonstrated here, many taxa in agricultural food webs
contribute to biological control of insects and weeds. In turn, as
researchers and practitioners in diversified farming systems, we
must focus more on integrating biological control of weeds and
We thank W.E. Snyder for useful discussions that helped the
manuscript. This project was supported by a USDA AFRI Fellowship
Grant (WNW-2010-05123) to DC.
References
Barton, B.T., Schmitz, O.J., 2009. Experimental warming transforms multiple
predator effects in a grassland food web. Ecol. Lett. 12, 1317–1325.
Bengtsson, J., Ahnstrom, J., Weibull, A.-C., 2005. The effects of organic agriculture on
biodiversity and abundance: a meta-analysis. J. Appl. Ecol. 42, 261–269.
Benton, T.G., Vickery, J.A., Wilson, J.D., 2003. Farmland biodiversity: is habitat
heterogeneity the key? Trends Ecol. Evol. 4, 182–188.
Bianchi, F.J.J.A., Booij, C.J.H., Tscharntke, T., 2006. Sustainable pest regulation in
agricultural landscapes: a review on landscape composition, biodiversity, and
natural pest control. Proc. R. Soc. B 273, 1715–1727.
Bracken, M.S., Friberg, S.E., Gonzalez-Dorantes, C.A., Williams, S.L., 2008. Functional
consequences of realistic biodiversity changes in a marine ecosystem. Proc.
Natl. Acad. Sci. USA 105, 924–928.
Byrnes, J.E., Stachowicz, J.J., 2009. The consequences of consumer diversity loss:
different answers from different experimental designs. Ecology 90, 2879–2888.
16
D.W. Crowder, R. Jabbour / Biological Control 75 (2014) 8–17
Cardinale, B.J., Duffy, J.E., Gonzalez, A., Hooper, D.U., Perrings, C., Venail, P., Narwani,
A., Mace, G.M., Tilman, D., Wardle, D.A., Kinzig, A.P., Daily, G.C., Loreau, M.,
Grace, J.B., Larigauderie, A., Srivastava, D.S., Naeem, S., 2012. Biodiversity loss
and its impact on humanity. Nature 486, 59–67.
Cardinale, B.J., Harvey, C.T., Gross, K., Ives, A.R., 2003. Biodiversity and biocontrol:
emergent impacts of a multi-enemy assemblage on pest suppression and crop
yield in an agroecosystems. Ecol. Lett. 9, 857–865.
Cardinale, B.J., Srivastava, D.S., Duffy, J.E., Wright, J.P., Downing, A.L., Sankaran, M.,
Jouseau, C., 2006. Effects of biodiversity on the functioning of trophic groups
and ecosystems. Nature 443, 989–992.
Casula, P., Wilby, A., Thomas, M.B., 2006. Understanding biodiversity effects on prey
in multi-enemy systems. Ecol. Lett. 9, 995–1004.
Chisholm, P., Gardiner, M., Moon, E., Crowder, D.W., isholm et al. 2014. Exploring
the toolbox for investigating impacts of habitat complexity on biological
control. Biol. Control 75, 48–57.
Cook-Patton, S.C., McArt, S.H., Parachnowitsch, A.L., Thaler, J.S., Agrawal, A.A., 2011.
A direct comparison of the consequences of plant genotypic and species
diversity on communities and ecosystem function. Ecology 92, 915–923.
Connolly, J., 1988. What is wrong with replacement series. Trends Ecol. Evol. 3, 24–
26.
Crowder, D.W., Northfield, T.D., Gomulkiewicz, R., Snyder, W.E., 2012. Conserving
and promoting evenness: organic farming and fire-based wildland management
as case studies. Ecology 93, 2001–2007.
Crowder, D.W., Northfield, T.D., Strand, M.R., Snyder, W.E., 2010. Organic agriculture
promotes evenness and natural pest control. Nature 466, 109–112.
Crutsinger, G.M., Collins, M.D., Fordyce, J.A., Gompert, Z., Nice, C.C., Sanders, N.J.,
2006. Plant genotypic diversity predicts community structure and governs an
ecosystem process. Science 313, 966–968.
Davis, A.S., Hill, J.D., Chase, C.A., Johanns, A.M., Liebman, M., 2012. Increasing
cropping system diversity balances productivity, profitability and
environmental health. PloS ONE 7, e47149.
Davis, A.S., Raghu, S., 2010. Weighing abiotic and biotic influences on weed seed
predation. Weed Res. 50, 402–412.
Donald, P.F., Green, R.E., Heath, M.F., 2001. Agricultural intensification and the
collapse of Europe’s farmland bird populations. Proc. Roy. Soc. B 268, 25–29.
Evans, D.M., Pocock, M.J.O., Brooks, J., Memmott, J., 2011. Seeds in farmland foodwebs: Resource importance, distribution and the impacts of farm management.
Biol. Conserv. 144, 2941–2950.
Finke, D.L., Denno, R.F., 2004. Predator diversity dampens trophic cascades. Nature
429, 407–410.
Finke, D.L., Snyder, W.E., 2008. Niche partitioning increases resource exploitation by
diverse communities. Science 321, 1488–1490.
Finke, D.L., Denno, R.F., 2005. Predator diversity and the functioning of ecosystems:
the role of intraguild predation in dampening trophic cascades. Ecol. Lett. 8,
1299–1306.
Foley, J.A., Ramankutty, N., Brauman, K.A., Cassidy, E.S., Gerber, J.S., Johnston, M.,
Mueller, N.D., O’Connell, C., Ray, D.K., West, P.C., Balzer, C., Bennett, E.M.,
Carpenter, S.R., Hill, J., Monfreda, C., Polasky, S., Rockstrom, J., Sheehan, J.,
Siebert, S., Tilman, D., Zaks, D.P.M., 2011. Solutions for a cultivated planet.
Nature 478, 337–342.
Food and Agricultural Organization of the United Nations (FAO). 2013. FAOSATS:
Crop yields. Available online: http://faostat.fao.org.
Frank, S.D., Shrewsbury, P.M., Denno, R.F., 2010. Effects of alternative food on
cannibalism and herbivore suppression by carabid larvae. Ecol. Entomol. 35,
61–68.
Frank, S.D., Shrewsbury, P.M., Denno, R.F., 2011. Plant versus prey resources:
Influence on omnivore behavior and herbivore suppression. Biol. Control 57,
229–235.
Franke, A.C., Lotz, L.A.P., Van Der Burg, W.J., Van Overbeek, L., 2009. The role of
arable weed seeds for agroecosystem functioning. Weed Res. 49, 131–141.
Gable, J., Crowder, D.W., Northfield, T., Steffan, S., Snyder, W.E., 2012. Niche
engineering reveals complementary resource use. Ecology 93, 1994–2000.
Gabriel, D., Roschewitz, I., Tscharntke, T., Thies, C., 2006. Beta diversity at different
spatial scales: plant communities in organic and conventional agriculture. Ecol.
Appl. 16, 2011–2021.
Gaines, H.R., Gratton, C., 2010. Seed predation increases with ground beetle
diversity in a Wisconsin (USA) potato agroecosystem. Agric. Ecosyst. Environ.
137, 329–336.
Gibbons, D.W., Bohan, D.A., Rother, P., Stuart, R.C., Haughto, A.J., Scott, R.J., Wilson,
J.D., Perry, J.N., Clark, S.J., Dawson, R.J.G., Firbank, L.G., 2006. Weed seed
resources for birds in fields with contrasting conventional and genetically
modified herbicide-tolerant crops. Proc. R. Soc. B 273, 1921–1928.
Goosey, H.B., Hatfield, P.G., Lenssen, A.W., Blodgett, S.L., Kott, R.W., 2005. The
potential role of sheep in dryland grain production systems. Agric. Ecosyst.
Environ. 111, 349–353.
Goosey, Hayes.B., 2012. A degree-day model of sheep grazing influence on alfalfa
weevil and crop characteristics. J. Econ. Entomol. 105, 102–112.
Grez, A.A., Gonzalez, R.H., 1995. Resource concentration hypothesis: effect of host
plant patch size on density of herbivorous insects. Oecologia 103, 471–474.
Gross, K., Rosenheim, J.A., 2011. Quantifying secondary pest outbreaks in cotton and
their monetary cost with causal-inference statistics. Ecol. Appl. 7, 2770–2780.
Hawes, C., Squire, G.R., Hallett, P.D., Watson, C.A., Young, M., 2010. Arable plant
communities as indicators of farming practice. Agric. Ecosyst. Environ. 138, 17–
26.
Hawkins, B.A., Mills, N.J., Jervis, M.A., Price, P.W., 1999. Is the biological control of
insects a natural phenomenon? Oikos 86, 493–506.
Hairston, N.G., Smith, F.E., Slobodkin, L.B., 1960. Community structure, population
control, and competition. Am. Nat. 94, 421–425.
Hillebrand, H., Bennett, D.M., Cadotte, M.W., 2008. Consequences of dominance: a
review of evenness effects on local and regional ecosystem processes. Ecology
89, 1510–1520.
Hillebrand, H., Matthiessen, B., 2009. Biodiversity in a complex world: consolidation
and progress in functional biodiversity research. Ecol. Lett. 12, 1405–1419.
Hole, D.G., Perkins, A.J., Wilson, J.D., Alexander, I.H., Grice, P.V., Evans, A.D., 2005.
Does organic farming benefit biodiversity? Biol. Conserv. 122, 113–130.
Hooper, D.U., Chapin III, F.S., Ewel, J.J., Hector, A., Inchausti, P., Lavorel, S., Lawton,
J.H., Lodge, D.M., Loreau, M., Naeem, S., Schmid, B., Setala, H., Symstad, A.J.,
Vandermeer, J., Wardle, D.A., 2005. Effects of biodiversity on ecosystem
functioning: a consensus of current knowledge. Ecol. Monogr. 7, 3–35.
Hopper, K.R., Roush, R.T., Powell, W., 1993. Management of genetics of biologicalcontrol introductions. Annu. Rev. Entomol. 38, 27–51.
Hughes, A.R., Inouye, B.D., Johnson, M.T.J., Underwood, N., Vellend, M., 2008.
Ecological consequences of genetic diversity. Ecol. Lett. 11, 609–623.
Huston, M.A., 1997. Hidden treatments in ecological experiments: re-evaluating the
ecosystem function of biodiversity. Oecologia 110, 449–460.
Inouye, B.D., 2001. Response surface experimental designs for investigating
interspecific competition. Ecology 82, 2696–2706.
Ives, A.R., Cardinale, B.J., Snyder, W.E., 2005. A synthesis of subdisciplines: predatorprey interactions, and biodiversity and ecosystem functioning. Ecol. Lett. 8,
102–116.
Jabbour, R., Crowder, D.W., Aultman, E.A., Snyder, W.E., 2011. Entomopathogen
biodiversity increases host mortality. Biol. Cont. 59, 277–283.
Jabbour, R., Zwickle, S., Gallandt, E.R., McPhee, K.E., Wilson, R.S., Doohan, D., 2013.
Mental models of organic weed management: Comparison of New England US
farmer and expert models. Renew. Agric. Food Syst.. http://dx.doi.org/10.1017/
S1742170513000185.
James, R.R., McEvoy, P.B., Cox, C.S., 1992. Combining the cinnabar moth (Tyria
jacobaeae) and the ragwort flea beetle (Longitarsus jacobaeae) for control of
ragwort (Senecio jacobaea): an experimental analysis. J. Appl. Ecol. 29, 589–596.
Johnson, M.T.J., Lajeunesse, M.J., Agrawal, A.A., 2006. Additive and interactive effects
of plant genotypic diversity on arthropod communities and plant fitness. Ecol.
Lett. 9, 24–34.
José-María, L., Sans, F.X., 2011. Weed seedbanks in arable fields: Effects of
management practices and surrounding landscape. Weed Res. 51, 631–640.
Kirk, D.A., Evenden, M.D., Mineau, P., 1996. Past and current attempts to evaluate
the role of birds as predators of insect pests in temperature agriculture. Curr.
Ornithol. 13, 175–269.
Kleijn, D., Baquero, R.A., Clough, Y., Diaz, M., Esteban, J.D., Fernandez, F., Gabriel, D.,
Herzog, F., Holzschuh, A., Johl, R., Knop, E., Kruess, A., Marshall, E.J.P., SteffanDewenter, I., Tscharntke, T., Verhulst, J., West, T.M., Yela, J.L., 2006. Mixed
biodiversity benefits of agri-environment schemes in five European counties.
Ecol. Lett. 9, 243–254.
Krebs, J.R., Wilson, J.D., Bradbury, R.B., Siriwardena, G.M., 1999. The second silent
spring? Nature 400, 611–612.
Letourneau, D.K., Jedlicka, J.A., Bothwell, S.G., Moreno, C.R., 2009. Effects of natural
enemy biodiversity on the suppression of arthropod herbivores in terrestrial
ecosystems. Annu. Rev. Ecol. Evol. Syst. 40, 573–592.
Liebman, Matt., Gallandt, E.R., 1997. Many little hammers: Ecological management
of crop-weed interactions. Ecol. Agric., 291–343.
Loreau, M., Hector, A., 2001. Partitioning selection and complementarity in
biodiversity experiments. Nature 412, 72–76.
Losey, J., Denno, R., 1998. Positive predator-prey interactions: enhanced predation
rates and synergistic suppression of aphid populations. Ecology 79, 2143–2152.
Losey, J.E., Vaughan, M., 2006. The economic value of ecological services provided
by insects. Bioscience 56, 311–323.
Macfadyen, S., Gibson, R., Polaszek, A., Morris, R.J., Craze, P.G., Planque, R.,
Symondson, W.O., Memmott, J., 2009. Do differences in food web structure
between organic and conventional farms affect the ecosystem service of pest
control? Ecol. Lett. 12, 229–238.
Matson, P.A., Parton, W.J., Power, A.G., Swift, M.J., 1997. Agricultural intensification
and ecosystem properties. Science 277, 504–509.
McEvoy, P., Cox, C., Coombs, E., 1991. Successful biological control of ragwort,
Senecio jacobaea, by introduced insects in Oregon. Ecol. Appl. 1, 430–442.
Meyer, W.B., Turner II, B.L., 1992. Human population growth and global land-use/
cover change. Annu. Rev. Ecol. Syst. 23, 39–61.
Miller, L.A., Surlykke, A., 2001. How some insects detect and avoid being eaten
by bats: tactics and countertactics of prey and predator. Bioscience 51, 570–
581.
Mt. Pleasant, J., Schlather, K.J., 1994. Incidence of weed seed in cow (Bos sp.) manure
and its importance as a weed source for cropland. Weed Sci. 8, 304–310.
Naeem, S., Li, S.B., 1997. Biodiversity enhances ecosystem reliability. Nature 390,
507–509.
Naeem, S., Wright, J.P., 2003. Disentangling biodiversity effects on ecosystem
functioning: deriving solutions to a seemingly insurmountable problem. Ecol.
Lett. 6, 567–579.
Navntoft, S., Wratten, S.D., Kristensen, K., Esbjerg, P., 2009. Weed seed predation in
organic and conventional fields. Biol. Control 49, 11–16.
Northfield, T.D., Crowder, D.W., Jabbour, R., Snyder, W.E. 2012. Natural enemy
functional identity, trait-mediated interactions and biological control, in:
Ohgushi, T., Schmitz, O., Holt, R.D (Eds.), Trait-Mediated Indirect Interactions:
Ecological and Evolutionary Perspectives. Cambridge University Press, New
York, pp. 450–465.
D.W. Crowder, R. Jabbour / Biological Control 75 (2014) 8–17
Northfield, T.D., Snyder, G.B., Ives, A.R., Snyder, W.E., 2010. Niche saturation reveals
resource partitioning among consumers. Ecol. Lett. 13, 338–348.
Onstad, D.W., Crowder, D.W., Isard, S.A., Levine, E., Spencer, J.L., Bledsoe, L.W.,
O’Neil, M.E., Eisley, J.B., Gray, M.E., Ratcliffe, S.A., DiFonzo, C.D., Edwards, C.R.,
2003. Does landscape diversity affect the spread of the rotation-resistant
western corn rootworm (Coleoptera: Chrysomelidae)? Environ. Entomol. 32,
992–1001.
Paini, D.R., Funderburk, J.E., Reitz, S.R., 2008. Competitive exclusion of a worldwide
invasive pest by a native. Quantifying competition between two phytophagous
insects on two host plant species. J. Anim. Ecol. 77, 184–190.
Parmenter, R.R., MacMahon, J.A., 1988. Factors influencing species composition and
population sizes in a ground beetle community (Carabidae): Predation by
rodents. Oikos 52, 350–356.
Perkins, A.J., Anderson, G., Wilson, J.D., 2007. Seed food preferences of granivorous
farmland passerines. Bird Study 54, 46–53.
Petchey, O.L., 2003. Integrating methods that investigate how complementarity
influences ecosystem functioning. Oikos 101, 323–330.
Pingali, P.L., 2012. Green revolution: impacts, limits, and the path ahead. Proc. Natl.
Acad. Sci. USA 102, 12302–12308.
Polis, G.A., 1999. Why are parts of the world green? Multiple factors control
productivity and the distribution of biomass. Oikos 86, 3–15.
Polis, G.A., Strong, D.R., 1996. Food web complexity and community dynamics. Am.
Nat. 147, 813–842.
Ramirez, R.A., Snyder, W.E., 2009. Scared sick? Predator-pathogen facilitation
enhances exploitation of a shared resource. Ecology 90, 2832–2839.
Root, R.B., 1973. Organization of a plant-arthropod association in simple and diverse
habitats: the fauna of collards (Brassica oleracea). Ecol. Mono. 43, 95–124.
Roschewitz, I., Gabriel, D., Tscharntke, T., Thies, C., 2005. The effects of landscape
complexity on arable weed species diversity in organic and conventional
farming. J. Appl. Ecol. 42, 873–882.
Rosenheim, J.A., 2007. Intraguild predation: new theoretical and empirical
perspectives. Ecology 88, 2679–2680.
Roubos, C.R., Rodriguez-Saona, C., Isaacs, R., ubos et al. 2014. Scale-dependent
impacts of insecticides on arthropod biological control. Biol. Control 75, 28–38.
Russelle, M.P., Entz, M.H., Franzluebbers, A.J., 2007. Reconsidering Integrated Crop–
Livestock Systems in North America. Agron. J. 99, 325–334.
Schmitz, O.J., 2007. Predator diversity and trophic interactions. Ecology 88, 2415–
2426.
Schmitz, O.J., 2008. Effects of predator hunting mode on grassland ecosystem
function. Science 319, 952–954.
Schmitz, O.J., Barton, B.T., 2014. Climate change effects on behavioral and
physiological ecology of predator-prey interactions: implications for
conservation biological control. Biol. Control 75, 87–96.
Schweitzer, J.A., Bailey, J.K., Hart, S.C., Whitham, T.G., 2005. Nonadditive effects of
mixing cottonwood genotypes on litter decomposition and nutrient dynamics.
Ecology 86, 2834–2840.
17
Sih, A., Englund, G., Wooster, D., 1998. Emergent impacts of multiple predators on
prey. Trends Ecol. Evol. 13, 350–355.
Snyder, W.E., Snyder, G.B., Finke, D.L., Straub, C.S., 2006. Predator biodiversity
strengthens herbivore suppression. Ecol. Lett. 9, 789–796.
Stephens, A.E.A., Srivastava, D.S., Myers, J.H., 2013. Strength in numbers? Effects of
multiple natural enemy species on plant performance. Proc. R. Soc. B 280,
20122756.
Stoate, C., Boatman, N.D., Borralho, R.J., Rio Carvalho, C., de Snoo, G.R., Eden, P., 2001.
Ecological impacts of arable intensification in Europe. J. Environ. Manag. 63,
337–365.
Strong, D.R., 1992. Are trophic cascades all wet? The redundant differentiation in
trophic architecture of high diversity ecosystems. Ecology 73, 747–754.
Tilman, D., 1999. Global environmental impacts of agricultural expansion: the need
for sustainable and efficient practices. Proc. Nat. Acad. Sci. USA 96, 5995–6000.
Trichard, A., Alignier, A., Biju-Duval, L., Petit, S., 2013. The relative effects of local
management and landscape context on weed seed predation and carabid
functional groups. Basic Appl. Ecol. 14, 235–245.
Turner, R.J., Davies, G., Moore, H., Grundy, A.C., Mead, A., 2007. Organic weed
management: A review of the current UK farmer perspective. Crop Prot. 26,
377–382.
Tylianakis, J.M., Tscharntke, T., Lewis, O.T., 2007. Habitat modification alters the
structure of tropical host–parasitoid food webs. Nature 445, 202–205.
Tylianakis, J.M., Binzer, A., 2014. Effects of global environmental changes on
parasitoid-host food webs and biological control. Biol. Control 75, 77–86.
Vanbergen, A.J., Woodcock, B.A., Koivula, M., Niemelä, J., Kotze, D.J., Bolger, T.,
Golden, V., Dubs, F., Boulanger, G., Serrano, J., Lencina, J.L., Serrano, A., Aguiar, C.,
Grandchamp, A.-C., Stofer, S., Szél, G., Ivits, E., Adler, P., Markus, J., Watt, A.D.,
2010. Trophic level modulates carabid beetle responses to habitat and
landscape structure: a pan-European study. Ecol. Entomol. 35, 226–235.
Welch, K.D., Harwood, J.D., 2014. Temporal dynamics of natural enemy-pest
interactions in a changing environment. Biol. Control 75, 18–27.
Williams, C.L., Liebman, M., Westerman, P.R., Borza, J., Sundberg, D., Danielson, B.,
2009. Over-winter predation of Abutilon theophrasti and Setaria faberi seeds in
arable land. Weed Res. 49, 439–447.
Wilson, R.S., Tucker, M.A., Hooker, N.H., LeJeune, J.T., Doohan, D., 2008. Perceptions
and beliefs about weed management: Perspectives of Ohio grain and produce
farmers. Weed Technol. 22, 339–350.
Wolfe, N.D., Dunavan, C.P., Diamond, J., 2007. Origins of major human infectious
diseases. Nature 447, 279–283.
Wright, J.P., Naeem, S., Hector, A., Lehman, C., Reich, P.B., Schmid, B., Tilman, D.,
2006. Conventional functional classification schemes underestimate the
relationship with ecosystem functioning. Ecol. Lett. 9, 111–120.
Yachi, S., Loreau, M., 1999. Biodiversity and ecosystem productivity in a fluctuating
environment: the insurance hypothesis. Proc. Natl. Acad. Sci. USA 96, 1463–
1468.