97
CHAPTER SIX Eco-Evolutionary Dynamics of Agricultural Networks: Implications for Sustainable Management Nicolas Loeuille ,,1 , Sébastien Barot { , Ewen Georgelin ,, Grigorios Kylafis ,, Claire Lavigne } *Laboratoire EcoEvo, UMR 7625, UPMC, Paris, France Laboratoire Ecologie des Populations et des Communaute ´s, USC INRA 2031, Paris, France { Laboratoire BIOEMCO, UMR 7618, IRD, Paris, France } Laboratoire Plantes et Syste `mes de culture Horticoles, UR1115, INRA, Avignon Cedex, France 1 Corresponding author: e-mail address: [email protected] Contents 1. Introduction 340 2. Within Field, Applying Evolutionary Perspectives to the Selection of Agricultural Species 344 2.1 General effects of domestication and selection 344 2.2 Beyond the one trait approach: Accounting for trade-offs 348 2.3 Adapting to local practices and conditions: The importance of diversity 351 2.4 Beyond individuals: The influence of selection processes on the community and ecosystem context 353 3. Disturbances Due to Agriculture: Implications for Eco-Evolutionary Dynamics Within Surrounding Ecosystems 360 3.1 Nutrient enrichment and its ecological and evolutionary consequences in agricultural landscapes 361 3.2 Chemical warfare in agricultural landscapes: The ecological and evolutionary consequences of pesticide use 365 3.3 Effects of altering species composition and relative abundance of species in agricultural landscapes 372 4. Accounting for Spatial Heterogeneities: Dispersal, Fragmentation, and Evolution in Agricultural Landscapes 376 4.1 Characteristics of agricultural landscapes, past, present and future 376 4.2 Consequences of agricultural landscape structure in terms of gene flow 380 4.3 Consequences of spatial modifications from a demographic point of view 385 4.4 Consequences of spatial structure for pairwise co-evolution 388 4.5 Beyond pairwise interactions: Consequences of eco-evolutionary dynamics for community structure and composition 390 4.6 Land sparing versus land sharing, from an evolutionary point of view 393 Advances in Ecological Research, Volume 49 # 2013 Elsevier Ltd ISSN 0065-2504 All rights reserved. http://dx.doi.org/10.1016/B978-0-12-420002-9.00006-8 339

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CHAPTER SIX

Eco-Evolutionary Dynamicsof Agricultural Networks:Implications for SustainableManagementNicolas Loeuille�,†,1, Sébastien Barot{, Ewen Georgelin�,†,Grigorios Kylafis�,†, Claire Lavigne}*Laboratoire EcoEvo, UMR 7625, UPMC, Paris, France†Laboratoire Ecologie des Populations et des Communautes, USC INRA 2031, Paris, France{Laboratoire BIOEMCO, UMR 7618, IRD, Paris, France}Laboratoire Plantes et Systemes de culture Horticoles, UR1115, INRA, Avignon Cedex, France1Corresponding author: e-mail address: [email protected]

Contents

1.

AdvISShttp

Introduction

ances in Ecological Research, Volume 49 # 2013 Elsevier LtdN 0065-2504 All rights reserved.://dx.doi.org/10.1016/B978-0-12-420002-9.00006-8

340

2. Within Field, Applying Evolutionary Perspectives to the Selection of Agricultural

Species

344 2.1 General effects of domestication and selection 344 2.2 Beyond the one trait approach: Accounting for trade-offs 348 2.3 Adapting to local practices and conditions: The importance of diversity 351 2.4 Beyond individuals: The influence of selection processes on the community

and ecosystem context

353 3. Disturbances Due to Agriculture: Implications for Eco-Evolutionary Dynamics

Within Surrounding Ecosystems

360 3.1 Nutrient enrichment and its ecological and evolutionary consequences in

agricultural landscapes

361 3.2 Chemical warfare in agricultural landscapes: The ecological and evolutionary

consequences of pesticide use

365 3.3 Effects of altering species composition and relative abundance of species in

agricultural landscapes

372 4. Accounting for Spatial Heterogeneities: Dispersal, Fragmentation, and Evolution

in Agricultural Landscapes

376 4.1 Characteristics of agricultural landscapes, past, present and future 376 4.2 Consequences of agricultural landscape structure in terms of gene flow 380 4.3 Consequences of spatial modifications from a demographic point of view 385 4.4 Consequences of spatial structure for pairwise co-evolution 388 4.5 Beyond pairwise interactions: Consequences of eco-evolutionary dynamics

for community structure and composition

390 4.6 Land sparing versus land sharing, from an evolutionary point of view 393

339

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340 Nicolas Loeuille et al.

5.

Perspectives and Challenges 395 Acknowledgements 398 Appendix A. Evolution of the investment into nutrient uptake, effects on emergentfunctioning 398

Appendix B. Mixing group selection and individual selection in co-evolutionarymodels

401

Appendix C. Effects of enrichment on the control of biomass within a tri-trophic foodchain when the herbivore evolves

402

Appendix D. Eco-evolutionary dynamics in a plant–herbivore–predator confronted toinsecticides

407

Appendix E. Evolution of specialization rate of the pest and its ecologicalconsequences

413

Glossary

416 References 417

Abstract

Community and ecosystem ecology are paying increasing attention to evolutionarydynamics, offering a means of attaining a more comprehensive understanding of eco-logical networks and more efficient and sustainable agroecosystems. Here, we reviewhow such approaches can be applied, and we provide theoretical models to illustratehow eco-evolutionary dynamics can profoundly change our understanding ofagricultural issues. We show that community evolution models can be used in severalcontexts: (1) to improve the selection of agricultural organisms within the context oftheir ecological networks; (2) to predict and manage the consequences of agriculturaldisturbances on the ecology and evolution of ecological networks; and (3) to designagricultural landscapes that benefit from network eco-evolutionary dynamics, but with-out negative impacts. Manipulation of landscape structure simultaneously affects bothcommunity ecological dynamics (e.g., by modifying dispersal and its demographiceffects) and co-evolution (e.g., by changing gene flows). Finally, we suggest that futuretheoretical developments in this field should consider appropriate co-evolutionarymodels and ecosystem services.

1. INTRODUCTION

Models and experiments in ecology have increasingly incorporated

evolutionary dynamics in the context of community structure and ecosys-

tem functioning (Agrawal et al., 2012; Brannstrom et al., 2012; Loeuille

and Loreau, 2009). Such models have been coined ‘community evolution

models’ (Brannstrom et al., 2012; Loeuille and Loreau, 2009). They link

evolutionary processes driven by individual fitness to emergent properties

of ecological networks, linking small-scale processes to large-scale patterns.

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341Eco-Evolutionary Dynamics in Agroecosystems

Evolutionary dynamics affect important structural attributes of ecological

communities, such as connectance or the number of trophic levels of a food

web, via the evolution of body size (Loeuille and Loreau, 2005) or adaptive

foraging (Beckerman et al., 2006). Similarly, the structure of mutualistic net-

works (e.g., plant–pollinator or plant–seed disperser) largely depends on the

co-evolution of partner species (Bascompte et al., 2006; Nuismer et al.,

2012). Evolution also modifies ecosystem functioning, in ways that are par-

ticularly relevant to agricultural systems: biomass, productivity, nutrient

cycling or ecosystem sustainability are all potentially affected. It can also pro-

duce counter-intuitive effects (Abrams and Matsuda, 2005) and the interplay

of ecological and evolutionary dynamics (hereafter, eco-evolutionary dynam-

ics) changes the distribution of nutrients within ecosystems (Boudsocq et al.,

2011), the way nutrients are recycled (De Mazancourt et al., 2001; Loeuille

et al., 2002) and their spatial distribution (Loeuille and Leibold, 2008a). Evo-

lution also constrains the resilience of ecological assemblages (Kondoh, 2003;

Loeuille, 2010a, b), thereby affecting the sustainability of the system.

Agricultural development is based on a selective process (Gepts, 2004), so

it is particularly important to incorporate rigorous evolutionary thinking

into it. Human populations have long chosen species and favoured traits that

increase the productivity of cultivated organisms or make their cultivation

easier. Because selective processes are at the heart of agricultural develop-

ments, it should be possible to implement current progresses linking trait

evolution with emergent properties of ecosystems, to improve selection

and the management of agricultural landscapes.

An immediate difficulty, however, results from differences in the ways

selective processes are considered in ecological versus agricultural contexts.

Because evolutionary concepts are often phrased differently in ecology and

agriculture, precise definitions of key concepts are necessary (see Glossary).

Selection in agriculture is driven by the choice, conscious or otherwise, of

cultivated species and of their characteristics (Meyer et al., 2012). Such a

choice is often based on total biomass or yield (Perales et al., 2004;

Vigouroux et al., 2011) and the process de facto relies on a group selection

criterion (Denison et al., 2003, but see Duputie et al., 2009). In contrast,

evolutionary ecology usually considers phenotypic changes based on indi-

vidual or gene fitness (Dieckmann and Law, 1996; Fisher, 1930;

Hamilton, 1964). Thus, fitness in community evolutionmodels can be com-

plex, due to the need to account for multispecies assemblages and interaction

networks (Ito and Ikegami, 2006; Loeuille and Loreau, 2005), as well as var-

ious disturbance scenarios (Loeuille and Loreau, 2004; Norberg et al., 2012).

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342 Nicolas Loeuille et al.

Here, we will discuss what agriculture can learn from these community

evolution models, as well as associated experimental and empirical works in

evolutionary ecology. Particularly, are there lessons to be learned to recon-

cile economic goals (e.g., an appropriate level of food production) with the

desire to conserve ecological networks? A few reviews have already shown

how evolutionary arguments can be applied to agricultural systems

(Denison, 2012; Denison et al., 2003; Thrall et al., 2011), by providing case

studies and examples of the many ways in which we can use evolutionary

ecology to design a better agriculture. Here, although some arguments will

be based on single-species evolution studies, we will focus on tackling the

links between evolutionary and community contexts, drawing extensively

on community evolution models and ecological theories.

We divide our central theme into two parts. (i) we ask how do commu-

nity evolution models help us to guide the choice of cultivated species, traits

or genes for a sustainable agriculture? This requires the assessment of how

these choices affect the co-evolution of species within associated ecological

networks and their consequences in terms of yield and sustainability. (ii) we

ask, can we use community evolution models to predict the consequences of

disturbances due to agriculture on surrounding ecosystems? Such distur-

bances include fertilization, pesticide use, changes in species community

composition and landscape modifications. Empirical observations suggest

that each of these disturbances has far reaching implications for the ecology

and evolution of the natural communities that abut agricultural systems

(Klein et al., 2007; Tscharntke et al., 2005). We also ask whether current

ecological and evolutionary theory can provide guidelines to limit the con-

sequences of these disturbances and preserve ecological networks.

It is necessary to point out that agricultural systems vary considerably in

how they create heterogeneities and disturbances (Malezieux, 2011; Massol

and Petit, 2013, Chapter 5 of this volume). From a low-intensity gatherer

mode of farming to intensive industrialized agriculture, a continuum exists

along which disturbances increase. Malezieux (2011) distinguishes ‘tradi-

tional agriculture’ from ‘intensive agriculture’ and warns that much of the

surface currently occupied by agriculture is not, despite common miscon-

ceptions, intensively managed. He also proposes to manage agriculture

through ‘ecological intensification’, relying on services provided by ecolog-

ical networks (Dore et al., 2011; Malezieux, 2011). Here, we will try to state

for each example whether it is applicable to intensive or traditional agricul-

ture, within agricultural fields or in surrounding ecosystems.

Justifications for questions (i) and (ii) can take several forms. One

important motivation is the conservation of species diversity. Agricultural

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343Eco-Evolutionary Dynamics in Agroecosystems

development has triggered marked declines in species diversity (Robinson

and Sutherland, 2002), although some species, such as those associated

with downland, heathlands and other semi-natural or farmed landscapes,

have benefited (Eriksson, 2012). The maintenance of (at least some)

diversity is important, not only for the intrinsic value of species, but also

because it generally enhances ecosystem functioning (Loreau et al., 2001),

and another important justification emerges from this point. When eco-

systems within the field and in the surrounding landscape are functioning

effectively, human populations receive important economical benefits,

commonly called ‘ecosystem services’ (Costanza et al., 1997; Raffaelli

and White, 2013; Bohan et al., 2013, Chapter 1 of this volume). Sev-

eral of these ecosystem services directly benefit agriculture, such as pol-

lination (Klein et al., 2007; Tscharntke et al., 2005), biological control of

pests by pathogens and predators (Crowder et al., 2010; Macfadyen et al.,

2011; Thrall et al., 2011) and the efficient recycling of nutrient and main-

tenance of soil fertility (Young and Crawford, 2004). Community assem-

bly and co-evolution have built the ecological networks that provide

these services, so community evolution models should be able to explain

the impact of agriculture on them.

We organize the text in three different parts, focusing on evolutionary

ecology within fields, around them, and then across the whole agricul-

tural landscape. In the first part, within agricultural fields, we ask how

evolutionary ecology in general and community evolution models in par-

ticular can guide the choice of cultivated species and of their traits, to

account simultaneously for agricultural yield and ecological sustainability.

In the second part, concerning adjacent ecosystems, we propose that

community evolution models can be used to discuss the consequences

of agricultural disturbances on the ecology and evolution of natural net-

works. The third part, integrating agricultural fields and their adjacent

ecosystems, investigates how spatial evolutionary models may guide the

landscape management of agricultural systems. In each part, we build

the arguments by increasing levels of organizational complexity, moving

from a species perspective (often the cultivated species), to pairwise inter-

actions, then to community structure and finally to ecosystem function-

ing. A table summarizing main arguments and key references introduces

each part. At several junctures, we illustrate our statements by developing

new models tightly focused on a key question. The main result of each

model is then presented in a box, while detailed hypotheses and analyses

are provided in appendices. In the fourth and final part, we finish by indi-

cating possible future developments.

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344 Nicolas Loeuille et al.

2. WITHIN FIELD, APPLYING EVOLUTIONARYPERSPECTIVES TO THE SELECTION OF AGRICULTURAL

SPECIES

2.1. General effects of domestication and selection

Domestication is the outcome of artificial selection that leads to increased

adaptation of plants and animals to cultivation by humans (Gepts, 2004).

The most detailed studies concern crop domestication (Gepts, 2004;

Harlan et al., 1973) and we will focus on this literature here as a complete

review is beyond the scope of the present chapter. Instead, we will focus on

the main phenotypic traits that have been selected via domestication then

show that most selected strategies would perform poorly in natural selection

settings. Inevitably, such a selection procedure requires extensive use of con-

tinued inputs to maintain the cultivated species and traits, posing a sustain-

ability problem.

About 2500 plant species have been domesticated (Meyer et al., 2012) in

an ongoing process that is rooted in the transition from hunter-gatherers to

settled agriculture during the Neolithic. The first steps of domestication

were probably unconscious: wild plants were simply translocated across

newly man-made environments, which altered selective pressures. How-

ever, further selection of varieties has probably been conscious for a very

long time. Artificial selection inspired much of Darwin’s theory of evolution

by natural selection (Darwin, 1859) and he later dedicated a whole book to

this subject (Darwin, 1868) after being struck by the strength and pace of

artificial selection and the obvious link it makes between present selective

pressures and subsequent evolution.

Many different traits are often selected in several domestication events,

defining a domestication syndrome (Gepts, 2004; Harlan, 1992; Meyer

et al., 2012; Purugganan and Fuller, 2009). In cereals, two types of traits

are essential: an ability to germinate deep in the soil in open, disturbed habitats

(e.g., large seeds, loss of dormancy) and traits that facilitate harvesting (e.g.,

non-shattering seeds, maturation synchrony, reduced branching, and resis-

tance to tilling). Domestication also involved the loss of secondary metabo-

lites, allowing seeds or fruits to taste better or to be more easily digested

(Meyer et al., 2012). Changes in the reproductive strategy are frequently

encountered, from outcrossing to selfing or from sexual to vegetative repro-

duction. Such changes made further selection easier, as they increase repro-

ductive isolation and facilitate the rapid fixation of selected traits.

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Take home messages and key references for section 2

SectionAgriculture and cropbreeding

Ecology and evolutionaryecology References

Section

2.1

Domestication and crop

breeding is a diversified

evolutionary process

impacting many plant

traits, favouring

easy cultivation and higher

yields

Domestication and artificial

selection have selected traits

that could not evolve

through natural selection,

due to differences in

selective regimes

Meyer et al.,

2012

Purugganan

and Fuller,

2009

Section

2.2

Crop breeding often

involves group selection

based on collective

properties of crops such as

yield. Such a selection has

favoured fast-growing crops,

incurring costs in terms of

competitive abilities

Existing allocation trade-offs

between growth rates and

many different traits (e.g.,

competitive ability,

defences) constrain the

evolution of these traits

Denison

et al., 2003

Section

2.3

Crop breeding has resulted

in some homogenization of

crop plants associated with

the development of

agricultural practices

allowing homogeneous

environmental conditions

through high inputs. An

unsettled controversy exists

regarding the degree of

desired local adaptation of

crops to different

agricultural practices.

Modern crop breeding

has led to a drastic reduction

in within field genetic

diversity

Natural selection selects

diversified strategies as soon

as environment is not totally

homogeneous.

Ecology increasingly views

genetic diversity as an

important factor for

ecosystem functioning

Thrall and

Burdon,

2003

Murphy

et al., 2007

Van Bueren

et al., 2008

Crutsinger

et al., 2006

Section

2.4

Results from community

evolution models suggest

that crop breeding could be

constrained by trade-offs

between yield and

sustainability

Allocation and ecological

trade-off are likely to result

in trade-offs at the

community or ecosystem

scale, including

between different

ecosystem services

Fussman

et al., 2007

Strauss et al.,

2002

Box 6.1

Box 6.2

345Eco-Evolutionary Dynamics in Agroecosystems

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346 Nicolas Loeuille et al.

The selective process changes widely among agricultural systems: in tra-

ditional cropping systems, farmers manage local landraces based on networks

of seed exchanges (Pautasso et al., 2013), whereas since the end of the nine-

teenth century, selection has been increasingly undertaken by specialized

farmers, public organisms and private companies. This specialization of roles

has allowed the development of sophisticatedmethods using the latest devel-

opments of biological sciences: genetics, molecular markers and biotechnol-

ogies, and the development of a global multinational agroindustry (Dennis

et al., 2008; Pennisi, 2008). Such modern changes in selection methods are

intertwined with the development of modern agriculture based on mecha-

nization, and the use of fertilizers and pesticides.

Crop breeding has targeted varieties with higher yields and three major

staple crops—wheat, rice, and corn—provide about 45% of human-ingested

calories worldwide. The grain yield of these cereals doubled during the

twentieth century, modern crop breeding methods playing a key role in this

increase (Richards, 2000). A trait related to yield, the harvest index (ratio of

the grain biomass to the whole aerial biomass), has also doubled, reaching

values as high as 0.5 in most cereals (Hay, 1995; Richards, 2000). Increases

in the harvest index have been obtained by selecting increased allocation to

reproductive parts and decreased allocation to leaves and stems. The devel-

opment of dwarfed varieties during the Green Revolution has led to shorter

and stiffer stems that are less sensitive to falling over, or lodging (caused by

bending or breakage of the stem or root problems), even at high fertilization

rates. Prior to the 1970s, these trends were mostly driven by ‘defect elimi-

nation’ and ‘selection for yield’ (Donald, 1968). The first corresponds to

removing major flaws, such as weak straws for cereals or fragile skins for

tomatoes. The second is based on the need to increase production, regardless

of the plant traits involved (Donald, 1968).

Harvest indices have seemingly plateaud (Hay, 1995), and crop breeding

now targets traits that appear most promising to increase yield: resistance to

pests and pathogens and to unfavourable conditions (Witcombe et al., 2008),

such as drought (Pennisi, 2008) or saline soils (Richards, 1992). This is cru-

cial for exploiting new, less favourable, land and to cope with global changes

(e.g., climate change, soil degradation). Other selected traits focus on the

efficient acquisition and recycling of resources. Such research aimed at opti-

mizing root systems (De Dorlodot et al., 2007) and at increasing their capac-

ity to take up phosphorus (Gahoonia and Nielsen, 2004). Artificial selection

of nutrient use also goes beyond individual plant traits to consider the com-

plex relationships with soil microbial communities. Nutrient management is

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347Eco-Evolutionary Dynamics in Agroecosystems

a key point for future agricultural sustainability because: (1) mineral fertil-

izers are not produced sustainably (phosphorus is dug in mines, nitrogen

is industrially fixed using fossil fuels as an energy source); (2) fertilizers leak

from agroecosystems, altering functioning and contributing to greenhouse

gas emissions (e.g., N2O) (Vitousek et al., 1997).

Artificial selection could also aim to increase photosynthetic rates (Long

et al., 2006; Richards, 2000), but when expressed per unit land area these

have mostly increased through inputs (irrigation, inorganic fertilizer).

Whether higher photosynthetic rates through crop breeding are possible

is still hotly debated (Denison et al., 2003), and it would require manipula-

tion of many genes simultaneously, while most modern crop breeding usu-

ally targets just a few genes.

Domestication and artificial selection alter the composition of commu-

nities at all spatial scales: out of 2500 domesticated species, 20 major crops

cover about 12% of the land surface (Leff, 2004), while the worldwide num-

ber of seed plant species is estimated at about 260,000. About 22% of con-

tinental surfaces are used as pastures and rangelands planted to some extent

with domesticated grasses (Glemin and Bataillon, 2009). In addition to

reducing species diversity, modern agriculture and crop breeding have

developed a restricted number of pure or hybrid lines so that genetic diver-

sity is drastically reduced, both locally and globally (Haudry et al., 2007).

Reduced genetic diversity can happen early in the domestication due to bot-

tlenecks, with early farmers using a limited number of plant individuals as

progenitors (Reif et al., 2005). A second step in the genetic erosion occurred

when pure or hybrid lines were preferred to local landraces. Genetic diver-

sity in bread wheat cultivated in a region of France, for instance, has halved

since 1878 (Bonneuil et al., 2012). On the contrary, farmers in traditional

systems often increase genetic diversity through the creation of local land-

races adapted to different local uses, tastes, agricultural practices, and envi-

ronments (Elias et al., 2001; Pautasso et al., 2013).

Can evolutionary ecology shed new light on the whole process of

domestication and crop breeding and can it improve crop breeding and

the sustainability of agriculture? Evolutionary arguments are not always pre-

sent in the recent crop breeding literature (but see Harlan et al., 1973). For

example, whole review papers on crop breeding do not use the term ‘evo-

lution’ in its Darwinian sense (Good et al., 2004; Zhang, 2007). Only some

relatively recent initiatives have appealed to more evolutionary-oriented

thinking in agriculture and crop breeding (Denison, 2012; Denison et al.,

2003; Gepts, 2004; Thrall et al., 2011).

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348 Nicolas Loeuille et al.

Artificial selection has led to species and varieties that depend on humans

to reproduce, as they have lost the suitable reproduction and dispersal strate-

gies, and to grow, with the need for nutrient addition and physical protection.

It tends to produce organisms whose traits would usually be disfavoured by

natural selection. The loss of traits allowing independent reproduction is only

possible because artificial selection is based on group level criteria, such as yield

(Denison et al., 2003). Such traits incur low individual fitness and would be

expected to disappear through natural selection. For instance, non-shattering

stems would be disfavoured in nature, because seed would not fall to the gro-

und, but are systematically selected in harvesting systems where humans want

to collect the seed before the seed is lost (Harlan et al., 1973). In some cases,

improving crop varieties has reversed the effects of past natural selection.

Thinking of natural selection therefore allows an a priori assessment of future

crop breeding avenues: Denison et al. (2003) hypothesize that natural selec-

tion is likely to have already optimized the major physiological pathways of

plants (e.g., photosynthesis), so significant further improvement through crop

breeding is unlikely.

Because the units of selection are different in artificial and natural selec-

tion, most evolutionary models developed in plant ecology have no straight-

forward links with crop breeding. Some theoretical models study the

evolution of traits relevant to crops, including growth rate, nutrient uptake

efficiency (Boudsocq et al., 2011), defences against herbivores (Loeuille and

Leibold, 2008a), seed size (Geritz et al., 1999), plant height (Falster and

Westoby, 2003), and shoot–root ratio (Vincent and Vincent, 1996). How-

ever, many crop traits are entirely anthropocentric and essentially irrelevant

in natural systems, such as those linked to the taste of the food or allowing an

easier harvest (e.g., non-shattering grains).

2.2. Beyond the one trait approach: Accounting for trade-offsWhile we have listed above a list of traits targeted by artificial selection, they

should not be considered in isolation. Trade-offs may cause negative corre-

lations between two or more traits, so that considering just one does not give

an adequate description of selection dynamics. This concept is at the very

heart of most evolutionary models and the influence of trade-offs has been

systematically investigated in this context (De Mazancourt and Dieckmann,

2004; Loeuille and Loreau, 2004). Trade-offs also exist between traits rele-

vant to artificial selection and their analysis is critical in terms of agricultural

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349Eco-Evolutionary Dynamics in Agroecosystems

management, although they are often overlooked in such a context

(Denison, 2012; but see Thrall, 2013).

Trade-offs can emerge due to allocation constraints of resources, time, or

space. For instance, resources allocated to growth cannot be used for repro-

duction or to produce defensive compounds against herbivores or pathogens

(Herms and Mattson, 1992). In the same vein, opening stomata allows pho-

tosynthesis but increases losses of water by evapotranspiration.We list below

the trade-offs that are commonly considered in evolutionary ecology and

that are relevant to agroecosystems.

In ecology, selection for yield could be related to the r–K theory (Pianka,

1970), which can emerge from existing trade-offs between individual traits

(Rueffler et al., 2006a) High yield, fast growth rates (r strategies) are thus

traded-off against traits that improve competitive ability (K strategies), such

as large size, long generation time, so that crop breeding is likely to produce

poor competitors (Denison et al., 2003; Donald, 1968; Weiner, 2004). Crop

individuals have higher yield because they spare the resource otherwise

required to interfere efficiently with other individuals. If only conspecifics

are present, having poor-competitor r strategists in turn decreases competition,

andyield is further increased (Donald,1968). It also requires a release fromcom-

petitionwithwildplants.This typeof selection is thereforedependenton tillage

and herbicides. The loss of competitiveness favours short stems (dwarfed vari-

eties), upright and erect leaves and lower individual leaf surface area, whereas

natural selection usually selects against these traits.

Recognition of the growth/competition trade-off raises the question of

the optimal strategy for achieving more sustainable agriculture. In a succes-

sion context, fast-growing r strategies are progressively replaced by K strat-

egies. In agriculture, artificial selection maintains r strategies in the field.

From an evolutionary and ecosystem ecology point of view, this mainte-

nance at early stages of succession cannot be a sustainable strategy for two

reasons. First, this situation is far from equilibrium, so that the system is sus-

ceptible to abrupt changes, low resilience and invasions (Odum, 1953). Sec-

ond, this maintenance of r strategies requires large inputs. Lower-yielding

phenotypes would allow lower rates of fertilization, as suggested by a the-

oretical model (Zhang et al., 1999) and experiments (Gersani et al.,

2001). They could also tolerate cultivation on marginal soils or under

stressed conditions.

Growth rate is also often negatively correlated with defences against her-

bivores or pathogens, as resources allocated to growth cannot be used for

both purposes (Herms and Mattson, 1992). Also, the allocation of mineral

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350 Nicolas Loeuille et al.

nutrient to roots or leaves should attract pathogens and herbivores (Mattson,

1980; White, 2005). Most often, in nature, intermediate strategies along the

growth/defence trade-off are selected (De Mazancourt et al., 2001; Loeuille

and Leibold, 2008a; Loeuille and Loreau, 2004; Loeuille et al., 2002). Crops

selected on yields in the absence of pests are again at one extreme of the con-

tinuum. They clearly depend on inputs (pesticides, fungicides) to compen-

sate, again raising sustainability issues.

Trade-offs also limit the improvements that are possible through crop

breeding, suggesting great difficulties in enhancing photosynthetic rate

(Denison et al., 2003). Without accounting for trade-offs, some scientists

have proposed the breeding of a Green Super Rice (Zhang, 2007), with

increased: (1) resistance to herbivores, (2) resistance to drought, (3) nutrient

use efficiency, (4) grain quality, and (5) yield. Most of these traits are likely to

be linked via antagonistic constraints, for instance due to resource allocation.

As a more general illustration of the lack of consideration given to these con-

straints, we have searched ISI Web of Science using the terms ‘crop breed-

ing’ and ‘trade-off’ (in keywords, title, and abstract) and only found

27 articles. Ideally, for any selection procedure, existing trade-offs among

targeted traits should be assessed before deciding which combination is more

desirable, and indeed feasible, using a hierarchical set of criteria (yield, crop

quality, sustainability of the production, etc., e.g., Quilot-Turion

et al., 2012).

In extremecases, targetingone trait can lead to the loss of others (Ellers et al.,

2012; Ostrowski et al., 2007), and reduced genetic diversity in crop plants can

increase the likelihood of such losses. Examples of traits that are threatened by

modern crop breeding include: (1) the ability to interact with mycorrhizae

(Zhu et al., 2001), (2) symbiotic nitrogen fixation (Kiers et al., 2007), (3) inhi-

bition of nitrification (Subbarao et al., 2006), and (4) the ability to benefit from

earthworm activities (Noguera et al., 2011). These four examples suggest that

trade-off choices have consequences for interspecific interactions, and, by

extension, ecological networks. Artificial selection of higher growth rates

may therefore have a general effect of decreasing benefits from interaction-

related traits. In natural ecosystems, plants rely heavily on such traits, which

evolved in complex four-dimensional settings involving soil organic matter

and interactions with micro- and macro-organisms (Puga-Freitas et al.,

2012; Shahzad et al., 2012). Selection of such traits can play a critical role for

amore sustainablemanagementof agriculture, as once interaction traits arepre-

sent, ecologicalprocesses canpartly replace artificial inputs,viaecological inten-

sification (Dore et al., 2011; Malezieux, 2011).

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351Eco-Evolutionary Dynamics in Agroecosystems

Many trade-offs have already been documented in crop plants:

between seed size and seed number (Sadras, 2007), between nutrient

use efficiency and nutrient stress tolerance (Maia et al., 2011), between

water and phosphorus acquisition (Ho et al., 2005) or between perfor-

mance at low and high salinity levels (Richards, 1992). The growth/

defence trade-off seems to be widespread in wild plants (Lind et al.,

2013) and has been investigated in a number of crop plants (e.g., tomato:

Le Bot et al., 2009; sunflower: Mayrose et al., 2011), although no general

overview exists for the latter. There is an emerging consensus, however,

that fertilization affects herbivores (Butler et al., 2012) and pathogens

(Huber and Watson, 1974), as well as plant growth, in agricultural sys-

tems. This trade-off is thus critical as it may provide an important key

to conceive alternatives to pesticide use.

The potential trade-off between nutrient uptake, at high and low con-

centrations of fertilizer, is controversial both for wild plants and for crops

(Craine, 2009), and results to date have produced little or no supporting evi-

dence (Hasegawa, 2003; Reich et al., 2003), suggesting that the same crop

varieties may be cultivated regardless of the fertilization practice.

This assessment of costs, and of trade-offs in general, is central to evolu-

tionary ecology and for agronomy. In ecology, the results of community evo-

lution models depend mostly on the shape of trade-offs, but such information

is often unavailable (De Mazancourt and Dieckmann, 2004; Loeuille and

Loreau, 2004). Artificial selection in agriculture could provide very important

data to address this gap, creating a synergy between the two disciplines.

2.3. Adapting to local practices and conditions: Theimportance of diversity

Diverse conditions, in space or time, typically favour a diversification of

strategies through natural selection (Thrall and Burdon, 2003; Thompson,

1999), unless strong homogenizing are operating via high gene flow imped-

ing local specialization (Harrison and Hastings, 1996; Day, 2001; Loeuille

and Leibold, 2008b; Urban et al., 2008), or if generalists are equally fit as

specialists, regardless of environmental conditions. This last condition

implicitly assumes an absence of trade-off. These conditions are very strin-

gent, so that diversification due to environmental variation is expected to

happen in most natural conditions.

This contrasts markedly with the situation in intensive agricultural sys-

tems, with a few species and a low genetic diversity. From an evolutionary

point of view, this homogenization of cultivated plants can be interpreted as

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352 Nicolas Loeuille et al.

either we have indeed managed to select highly generalist no-cost varieties,

or the heterogeneity of conditions in which crops grow is insufficient to sup-

port many specialized varieties. The second explanation seems a better fit to

our understanding of agricultural systems, for several reasons. First, a tem-

poral correlation exists between diversity decline of cultivated plants and

homogenization of their growing conditions, which is obtained by tillage,

herbicides and pesticides that remove biotic interactions, while irrigation

and mineral fertilization smooth out variations in resource conditions.

A more sustainable agriculture would require management that changes

modern agricultural practices and crop breeding approaches simultaneously

(Tilman, 1999), but because both belong to the same technological regime,

this creates inertia that impedes the development of alternative strategies

(Vanloqueren and Baret, 2009).

A direct corollary is that more sustainable crop varieties should fit better

to their local environment and practices. Interactions between genotypes

and environmental conditions determine crop performance (Van Bueren

et al., 2008), and differences among varieties have been found for systems

where there is variation in intercropping practices (O’leary and Smith,

1999), mineral fertilization (Atlin and Frey, 1990; but see Hasegawa,

2003), or soil salinity (Kelman and Qualset, 1991). Yields of 35 genotypes

of soft white winter wheat were compared between organic and conven-

tional practices, and a significant interaction between the genotype and

cropping system was evident in four out of five locations (Murphy et al.,

2007). This suggests that varieties to be used in organic farming should be

selected for under the conditions of organic farming (Przystalski et al.,

2008), whereas at present it currently largely relies on ‘conventional’ vari-

eties. Breeding particular varieties for this type of agriculture could increase

yields and improve organic agriculture sustainability (Wolfe et al., 2008).

Since a better adaptation to local environments and practices should enhance

sustainability and yields, such a strategy should inevitably be favoured,

although the costs entailed by such local adaptation also need to be

accounted for, such as those related to economies of scale, labour, agronomic

management, and crop handling.

The loss of genetic diversity of cultivated organisms is detrimental to agri-

culture sustainability because it limits the scope of potential future evolution

(Dieckmann and Law, 1996; Lande, 1979). This, in addition to the fact that

farmers in intensive systems no longer produce their own seeds,

will compromise future local adaptation in crops. In traditional cropping sys-

tems, farmers adjust the landraces to suit their own farm. It may be argued that

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353Eco-Evolutionary Dynamics in Agroecosystems

modern tools of artificial selection (e.g., genetic engineering) can offset the

lack of on-farm local evolution, but local adaptation involves many, often

unknown, niche dimensions, and understanding its significance requires a pre-

dictive framework we still lack. In traditional agriculture, on the other hand,

on-farm selection and seed exchange networks allow the necessary evolution

of crops in the face of global change (Bellon et al., 2011) and the evolution of

resistance to pathogens (Paillard et al., 2000).

Mechanisms identified from ecological biodiversity-ecosystem function-

ing studies also suggest how diversity loss can be detrimental to agricultural

sustainability. In natural ecosystems, higher plant species diversity produces

higher biomass, primary production, as well as more resilient systems

(Loreau et al., 2001). Different species react to environmental variations,

often in uncorrelated ways, which decreases the overall variance of the com-

munity (Yachi and Loreau, 1999). This ‘portfolio’ mechanism, postulated

for species diversity, also applies to genetic diversity (Vellend and Geber,

2005). Genetic diversity within species also influences ecosystem function-

ing (Crutsinger et al., 2006; Kotowska et al., 2010), and in agroecosystems,

for instance, a mixture of genotypes with different resistance levels may

impede a pathogen outburst by reducing the overall risk of contamination

(Zhu et al., 2000; Bohan et al., 2013, Chapter 1 of this volume).

2.4. Beyond individuals: The influence of selection processeson the community and ecosystem context

Choices of domesticated species and artificial selection of their traits will

ultimately affect the overall structure of ecological networks. Indeed, the

selection imposed on cultivated organisms, often the dominant species

of their ecosystem, affects interspecific interaction in multiple ways. We

previously described trade-offs based on allocation costs, but while these

are undoubtedly important drivers of the evolution of some traits

(Herms and Mattson, 1992), the evolution of other traits incurs a different

type of cost. They may be constrained by the balance of various ecological

interactions (thereafter, ecological costs), rather than by energy allocation

within individuals.

Strauss et al. (2002) propose that anti-herbivore defences can have dif-

ferent allocation and/or ecological costs. Importantly, ecological costs

directly link trait selection to community aspects. For instance, Muller-

Scharer et al. (2004) suggest that although some phenolic compounds are

effective against generalist herbivores, they may attract specialists. In princi-

ple, a plant species can then be invasive simply because their specialists are

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354 Nicolas Loeuille et al.

absent, reducing the costs of these defences to zero. Evolution of defences

that incur ecological costs affect herbivore insect communities in complex

ways (Courtois, 2010), and they can also affect higher trophic levels or

mutualistic species. In tobacco plants, nicotine compounds, usually consid-

ered to be primarily selected as herbivore defences (Krieger et al., 1971), are

negatively linked to the pollinator reliance of the plant (Adler et al., 2012).

Defences also affect nectar quality, so that herbivore repulsion is traded-off

against pollinator attraction (Adler and Bronstein, 2004). Some volatile

compounds serve as a defence but they can also act as cues for herbivore

behaviours, leading to attraction to damaged plants for example, and signals

to insects at higher trophic levels and pollinators (Poelman et al., 2008; Van

Zandt and Agrawal, 2004; Xiao et al., 2012). Artificial selection can directly

modify defensive traits (Carriere et al., 2010) and domestication often

involves modification of palatability or digestibility of the cultivated organ-

ism (Meyer et al., 2012). This has indirectly selected against secondary met-

abolic compounds that are used for defence. How a modification of these

defences affects the surrounding ecological network will, in turn, depend

on the ecological costs associated with these traits.

The importance of ecological trade-offs goes beyond insect communi-

ties, as other taxa and soil maintenance and recycling processes are also

involved. Tannin production, again a classical defensive compound, slows

down litter recycling by affecting the soil microbial loop (Grime et al.,

1996;Whitham et al., 2003). Community evolutionmodels also suggest that

defences affect recycling processes via the herbivore loop (De Mazancourt

et al., 2001). Defences could decrease root association with mycorrhizae,

impairing the plant’s ability to take up mineral nutrients (de Roman

et al., 2011). Plant defensive traits may therefore be linked to the long-term

maintenance and recycling efficiency of soils, a feature that underpins the

sustainability of agricultural yields (Malezieux, 2011). More recently, an

experimental study has even shown that defences may be traded-off against

competitive ability (Agrawal et al., 2012), thereby coupling two traditionally

separate issues in agriculture: competition from weeds and control of pests.

Community evolution models illustrate ways in which evolution under

ecological costs can affect network properties. For instance, Loeuille and

Leibold (2008a) introduced a model in which two defensive strategies are

incorporated, one incurring an allocation cost and the other having an eco-

logical cost. Because of ecological costs, under high nutrient conditions,

evolution of defences within the food web decreases species diversity and

the network also becomes more compartmentalized, with well-separated

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355Eco-Evolutionary Dynamics in Agroecosystems

food chains. Evolution constrained by ecological costs is also usually less pos-

itive for community resilience than when it is constrained by allocation costs

(Loeuille, 2010a). While such models suggest that ecological costs can have

considerable implications for community structure and stability, empirical

data to determine their consequences in an agricultural context are scarce.

Recent observations on rice, however, indicate that such implications can

be far-reaching (Xiao et al., 2012), as artificial selection of anti-herbivore

defensive strategies may indeed modify the structure and interactions of

the complete insect community, including parasitoids and predators. The

direct implementation of defensive traits in the absence of an assessment

of ecological costs for the maintenance of associated ecological networks

should therefore be treated with caution.

Evolution also affects ecosystem functioning and services (Fussman et al.,

2007), and, as an example, selection for the high yield and fast-growing strat-

egies favoured in agriculture can lead to a ‘tragedy of the commons’, in

which evolution of acquisition of shared resources leads to the overex-

ploitation and eventual extinction of these resources (Hardin, 1968) with

negative consequences for the sustainability of the system.

Emergent properties of the system such as biomass or mineral nutrient

availability are affected by the evolution of such traits in complex ways

(Boudsocq et al., 2011; Loeuille et al., 2002). For example, in the model

by Boudsocq et al. (2011) (see also Box 6.1), plant evolution of nutrient

uptake may lead to three contrasting evolutionary outcomes depending

on costs: (1) an evolutionary equilibrium is reached, (2) evolution negatively

affects the population viability, and (3) accumulation of mineral nutrient,

eventually followed by a switch to another limiting nutrient or factor. From

an agricultural perspective, only option (1) is sustainable. Outcome (2)

shows how evolution of plant strategy can affect biodiversity maintenance,

while outcome (3) highlight how it can have important consequences for

ecosystem functioning (here, changes in the limiting nutrient).

Let us focus on the first case, evolutionary equilibrium, whereby natural

selection leads to intermediate rates of nutrient uptake at which the availabil-

ity of mineral nutrient is minimal and the selected strategy does not maxi-

mize primary productivity or plant biomass. Modifying the model by

Boudsocq et al. (2011) to account more closely for artificial selection

(Box 6.1) and using primary productivity (yield, a common target for crop

breeding) as a criterion for selection then leads to higher mineral nutrient

availability accompanied by a faster rate of loss from the system, which cre-

ates a sustainability cost.

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BOX 6.1 Natural selection, artificial selection, and thefunctioning of ecosystemsThe model (Boudsocq et al., 2011) is based on a simple representation of nutrientcycling between three compartments (primary producers, dead organic matter,and limiting mineral nutrient) in an open ecosystem (see the detail of model inAppendix A). Nutrient availability is strictly determined by the quantity of mineralnutrient. Nutrient losses occur from the mineral nutrient pool, due to diffusion. Topredict theevolutionof investment intonutrientuptake,weuse theadaptivedynam-ics framework (Dieckmann and Law, 1996; Geritz et al., 1999).We consider a trade-offbetween nutrient uptake and plant nutrient turnover due to root and leaf mortality.The existence of this trade-off is supported bymanymechanisms and observations.For example, a higher investment into the root system to increase nutrient uptakemay (1) decrease the allocation of resources to defences against herbivores(Herms and Mattson, 1992), (2) increase the mineral nutrient concentration in theplant biomass which stimulates litter decomposition (Endara and Coley, 2011), (3)require theproductionofmany thin rootswithahigh turnover (Eissenstatetal., 2000).

Three types of evolutionary dynamics are possible depending on the trade-off shape: (1) evolution of plant traits lead to an accumulation of nutrient so thatanother nutrient type eventually becomes limiting, (2) evolution leads to a run-away towards higher uptake of nutrient, which asymptotically leads to plantextinction, (3) an evolutionary equilibrium is reached with an intermediate invest-ment in nutrient uptake. Analytical computations show that the availability ofmineral nutrient always decreases along evolutionary dynamics. Neither the bio-mass of primary producers nor primary productivity are maximized at the evolu-tionary equilibrium (Boudsocq et al., 2011).

Focusing on cases that lead to an evolutionary equilibrium, we illustrate theseresults inFig. 6.1, that showshowtwocompartments (mineralnutrientpool, plantbio-mass) andprimaryproductivity changedependingon thevalueof theevolving trait s,the investment intonutrientuptake.Natural selectionyieldsan svalue(s�, vertical solidline) that minimizes mineral nutrient availability (N�, horizontal solid line), thus mini-mizing mineral nutrient losses happening through diffusion out of the system.

Can this model help predicting the result of the artificial selection in cropplants? A possible scenario is that crop breeding is managed tomaximize primaryproductivity (f0) and thus drives crop plants towards the s0 value (doted verticallines). Compared with the previous ‘natural selection’ scenario, such a selectionincreases mineral nutrient availability (N0, horizontal dotted line) and associatednutrient losses (Fig. 6.1A). Therefore, our results suggest that artificial selectionper se, even in the absence of fertilization, may modify nutrient diffusion in waysthat are negative for agricultural sustainability.

The three other panels (B, C, D) of Fig. 6.1 display the same type of results afteran increase in fertilization (panel B), an increase in plant biomass export to mimicharvesting (C), or an increase in both (D). These changes increase the effect ofartificial selection, further enhancing nutrient losses. This result is not trivial.

356 Nicolas Loeuille et al.

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BOX 6.1 Natural selection, artificial selection, and thefunctioning of ecosystems—cont'd

Fertilization and crop harvesting can intuitively be thought as factors increasingnutrient losses at the ecological scale, but crop breeding could have mitigatedsuch effects. We find the contrary: crop breeding further increases nutrient losses.Our results illustrate that the intensification of agriculture (increase in harvestindex and mineral fertilization) interacts with crop breeding to determine thesustainability of agriculture (here characterized by nutrient losses). This resultis useful since the mitigation of nutrient losses is generally viewed independentlyeither from the point of view of agricultural practices (e.g., Gardner andDrinkwater, 2009) or of crop breeding (e.g., Good et al., 2004).

0

0 0

00

00

0

0

0

F

F

F

F

¢ ¢

¢

¢

¢

¢

¢¢

A B

C D

Figure 6.1 Each panel describes variations in the nutrient stock of primary pro-ducer (V0), the stock of mineral nutrient (N0) and primary productivity (F0) at theirecological equilibriums as a function of the investment into nutrient uptake (s). Inthe displayed cases an evolutionary equilibrium is reached (vertical solid line, s�)and natural selection has been show to minimize N (solid horizontal cases, N�).Crop breeding can be supposed to maximize primary productivity (vertical dottedline (s0)). This value of the investment into nutrient uptake leads to an availability ofmineral nutrient N0 (doted horizontal line). In all cases crop breeding leads to ahigher availability of mineral nutrient (N0>N�). Panel A is based on the sameparameters as in the original article (fertilization, 6 kg nutrient ha�1 year�1; bio-mass exportation, 0.1 year�1)(see original Fig. 6.2 in Boudsocq et al., 2011). PanelB uses a higher fertilization rate (30 kg nutrient ha�1 year�1). Panel C uses a higherexportation rate of plant biomass (0.5 year�1). Panel D uses both a higher fertili-zation and a higher exportation rate.

357Eco-Evolutionary Dynamics in Agroecosystems

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358 Nicolas Loeuille et al.

Such results raise an intriguing possibility: evolutionary dynamics can

lead to negative relationships between ecosystem services (here primary pro-

duction versus nutrient conservation). While such negative relationships

have already been noted (Bennet et al., 2009), their fundamental causes as

well as their consequences for ecosystem management, ecological engineer-

ing and agriculture are mostly unknown.We suggest that they may naturally

emerge from artificial selection.

In agricultural networks, species co-evolve with the cultivated plant.We

now consider a system in which cultivated species would be selected

through artificial selection and a wild herbivore species would follow natural

selection based on individual fitness (Box 6.2). The initial community

BOX 6.2 Mixing group and individual selection to study theco-evolution of agroecosystem networksMost often, community evolutionmodels that tackle co-evolution use individual fit-ness to determine evolutionary dynamics. In the case of agriculture, to study theco-evolution of a cultivated plant species and of the species from its surroundingnetwork, one may want to base the evolutionary dynamics of the plant speciesonanartificial group selectioncriterion, and theevolutionarydynamicsofother spe-cies based on individual fitness. We know of only one mixed model of this kind(Fletcher and Doebeli, 2013). We illustrate some possible implications of this ideausing the model described fully in Loeuille et al. (2002), which studies theco-evolution between defence investment by the plant, and herbivore attack rate.The plant–herbivore interaction ismodelled using a Lotka–Volterra function,whoserate depends on the two traits. Plant defences incur a cost in terms of growth whileherbivore investment in consumption incurs a mortality cost. In its initial version,based on natural selection, the plant–herbivore co-evolution model leads to a setof strategies (plant defences, herbivore attack rate) that allow the coexistence ofplants and herbivores, although someoscillationsmay exist around the equilibriumstate (Fig. 6.2A). If plant selection now depends on total biomass, for instancebecause the farmer selects phenotypes producing higher yields, higher defencesare always favoured, and co-evolutionary dynamics are profoundly affected. Even-tually, such a co-evolution causes the extinction of the herbivore (evolutionarymur-der, Fig. 6.2B, seealsoAppendixB). This incurs aneconomicbenefit if theherbivore isa crop pest. On the other hand, the argument is fairly general, so that theco-evolutionary dynamics associated with artificial selection can also produceunwanted extinctions. Of course, this model is largely simplified compared toreal agricultural systems, in which herbivores have not disappeared (thoughthey are usually less abundant). In any case, it illustrates a mechanism throughwhich co-evolution community structure may be profoundly affected by artificialselection.

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BOX 6.2 Mixing group and individual selection to study theco-evolution of agroecosystem networks—cont'd

Figure 6.2 Plant herbivore co-evolution depending on the selection regime. Inpanel A, plant defences co-evolve with herbivore attack rate depending on individ-ual fitness. The evolutionary outcome (black filled circle) is at the intersection of theherbivore isocline (light grey) and of the plant isocline (black). Herbivore evolution-ary dynamics always converge towards the isocline (plain line). For the plant isocline,evolutionary dynamics can be convergent (plain line) or divergent (dotted line).When the isoclines intersect in the plain part, evolutionary dynamics will lead tothe evolutionary singularity. In the dotted part, dynamics may cycle aroundthe singularity. In all instances, the full system plant–herbivore survives (i.e.,co-evolution stays within existence boundaries, dashed curve). For more detailssee Loeuille et al. (2002). Panel B: Same model, but plant selection is based onbiomass production instead of individual fitness. Under such conditions, more def-ended morphs are always favoured (Appendix B). This creates an ever increasingtrait value for the plant that leads to the extinction of the herbivore (dashed curve),that is, an evolutionary murder (Dercole et al., 2006).

359Eco-Evolutionary Dynamics in Agroecosystems

co-evolution model, where both species evolved out of natural selection,

predicts the maintenance of the complete community. Such maintenance

is unlikely when plants are subjected to group selection, where higher

defences are favoured, because they allow a higher plant biomass by reducing

herbivory. Eventually, such high defences cause the extinction of the her-

bivore. Therefore, artificial selection regime can have important implica-

tions not only for evolutionary dynamics, but also for side-effects on the

community structure and conservation issues of non-targeted species.

Such results of mixing artificial and natural selection pose interesting new

questions. As evolution affects the system’s functioning, a possible target for

crop breeding would be not only the plant species or traits, but also some

part of its connected ecological network. A clear example is the experiment

described in Swenson et al. (2000), who grew Arabidopsis in pots whose soil

was chosen from an ecosystem selection perspective. At each generation,

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360 Nicolas Loeuille et al.

pots were inoculated with soil samples selected from the biomass of an Ara-

bidopsis individual of the previous generation. Group selection then involved

all organisms present in the inoculate soil based on plant yield. Individual

plant biomass increased rapidly through this selection process. From a the-

oretical point of view, such outcomes raise intriguing questions, such as

what sets of species of the network should we incorporate in this ecosystem

selection processes? How many species or functional groups should be

involved?What constrains the efficiency of the selection process? All of these

aspects remain largely unknown.

3. DISTURBANCES DUE TO AGRICULTURE:IMPLICATIONS FOR ECO-EVOLUTIONARY DYNAMICS

WITHIN SURROUNDING ECOSYSTEMS

Lessons from eco-evolutionary models can applied when considering

the effects of agricultural disturbances (Gould, 1991; Thrall et al., 2011;

Verbruggen and Kiers, 2010), which typically correspond to strong selective

pressures on wild organisms, so fast evolution should be expected when suit-

able variability exists for associated heritable traits. It is critical to account for

these eco-evolutionary dynamics, as they affect the ultimate survival of these

organisms in the landscape (Ferriere et al., 2004; Loeuille and Leibold,

2008b) and produce indirect effects that propagate through ecological net-

works. Alternatively, adaptation can also take place via plasticity. Phenotypic

plasticity may limit the evolution of genetic responses to disturbances, by

modulating individual fitness (Ghalambor et al., 2007; Hendry et al.,

2011). Also, the degree of plasticity is itself under selection and ultimately

depends on the disturbance regime (frequency, amplitude and predictabil-

ity). We first list a few of the disturbances associated with agriculture. This

is not meant to be an exhaustive or exclusive list. Rather, we focus on dis-

turbances whose implications may be assessed using community evolution

models.

As modern approaches of crop selection often rely on minimizing envi-

ronmental variation through a wide use of nutrient additions and a strong

control of enemies, these efforts to simplify the system correspond to strong

selective pressures on the ecological networks in which the cultivated

organism is embedded. There are three main disturbances, which we

will now focus on in turn: nutrient addition, pesticide use, and habitat

alteration.

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Take home messages and key references for section 3

SectionAgriculture and cropbreeding

Ecology and evolutionaryecology References

Section

3.1

Agriculture exerts strong

selective pressures on wild

animals and plants through

direct inputs and habitat

modification

Fast evolution is expected

for variable traits, given the

strong disturbances

Thrall et al.,

2011

Loeuille and

Leibold, 2008aCo-evolution creates

indirect effects that

propagate throughout the

network

Section

3.2

Fertilization creates global

nutrient enrichment

Nutrient enrichment

modifies top-down

controls within food chains

by altering the selection

regime of defences

Oksanen et al.,

1981

Herms and

Mattson, 1992

Evolutionary impacts of

enrichment on

communities modify

important ecosystem

services

Loeuille and

Loreau, 2004

Box 6.3

Section

3.3

Pesticides exert strong

selective pressures that

trigger the evolution of

resistances. This incurs

large economic costs

Extra-mortality

evolutionary models can be

used to discuss such

disturbances

Mallet, 1989

Abrams and

Matsuda, 2005

Box 6.4

Evolution of resistance

depends on the community

context

Side-effects exist on non-

targeted species

Section

3.4

The homogenization of

biotic and abiotic

conditions by agriculture

should mostly select for

specialized interactions

Evolution of specialization

depends on species

abundance distributions in

communities

Evolution of specialization

affects the propagation of

Poisot et al.,

2011

Dore et al.,

2011

Symondson

et al., 2002Evolution of specialization

361Eco-Evolutionary Dynamics in Agroecosystems

disturbances in the system Box 6.5affects pest management

and ecosystem services

3.1. Nutrient enrichment and its ecological and evolutionaryconsequences in agricultural landscapes

Most natural ecosystems being limited by either N, P or K (Boudsocq et al.,

2012; Vitousek and Howarth, 1991), and agroecosystems are no exception.

Fertilization mainly consists of these nutrients. They are transported in sur-

rounding ecosystems through abiotic factors (diffusion, transport by water)

or through biotic factors (e.g., metaecosystem effects: Loreau et al., 2003).

Nutrient additions profoundly change the diversity and composition of

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362 Nicolas Loeuille et al.

natural communities, in terrestrial and aquatic systems (Tilman, 1999; Xia and

Wan, 2008) and are one of the major players in current global changes

(Langley and Megonigal, 2010; Reich, 2009).

From an evolutionary point of view, nutrient additions have direct selec-

tive effects, by alleviating costs of life history or physiological traits that are

energetically expensive. They also generate indirect selective effects on traits

related to interspecific interactions, as they modify the composition and

structure of ecological networks. We will analyse first the direct effect (allo-

cation costs), and then turn our attention to assessing indirect effects (com-

munity composition and diffuse co-evolution).

To illustrate how the evolution of traits with allocation costs alters the

dynamics of communities, the best starting point is to study the qualitative

effects it has on food chains. Comparing the impact of nutrient enrichment

on food chain dynamics without and with evolution will enable us to assess

its role in community structure and composition.

If a food chain is supplied with increased resources, two possibilities exist.

If biomass at each trophic level is bottom-up controlled—meaning that

competition for resources is the major driver of ecological dynamics—then

biomasses at all trophic levels will increase, because enrichment relaxes com-

petition constraints. This type of pattern holds well in some ecosystems

(Chase, 2003; White, 2005). In other instances, when the biomass at a given

level is determined by predation exerted on it (top-down control), the bio-

mass at the top level of the food chain will increase during enrichment, as

well as all odd levels from the top-down, while the biomasses at other tro-

phic levels remain constant or decrease (Hairston et al., 1960; Oksanen et al.,

1981). Again, empirical patterns can be found to support such top-down

control predictions (Persson et al., 1992; Ripple and Betscha, 2012).

Figure 6.3 in Box 6.3 illustrates such an ecological impact of nutrient

enrichment. Of course, the picture is inevitably more complicated when

omnivory and generalist feeding, which are common features of most eco-

logical networks, blur trophic levels into continua, but the same general

principles apply.

Biological control of crop pests by natural enemies is one of the many

ecosystem services on which the development of a sustainable agriculture

relies (Costanza et al., 1997; Crowder et al., 2010; Dore et al., 2011;

Malezieux, 2011). Top-down control models therefore provide conditions

under which biological controls should operate. For instance, considering

herbivore as pests, nutrient enrichment would increase pest biomass when

the food chain has two levels (plant–herbivore) or an even number of levels

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BOX 6.3 Herbivore pest evolution under nutrient enrichment:

(e.g., plant–herbivore–predator–super predator) (Fig. 6.3). Otherwise, we

expect biological control to act, and herbivores to be maintained at constant

levels by their enemies during the enrichment process.

Unfortunately, evolutionary dynamics erode the conditions under

which this top-down control is expected to happen. Many phenotypic traits

constraining top-down controls have allocation costs, so that their selection

regime may change when energetic constraints are relaxed. Allocation costs

are for instance well described for some plant defences (Herms and Mattson,

1992; Strauss et al., 2002). If nutrient enrichment selects for more defended

plants, energy is less easily transmitted up the food chain, so that top-down

controls are decreased (Loeuille and Loreau, 2004). This is also true for other

traits, such as vigilance, where a possible cost is the foregone time and energy

from activities not related to resource acquisition (Illius and Fitzgibbon,

1994). Again, if nutrient enrichment makes the resource more abundant

or more nutritious, higher vigilance can be favoured (Armstrong, 1979;

Leibold, 1996), which decreases energy transmission through the food

chain. Similar arguments could be made for the evolution of time par-

titioning between productive and unproductive habitat (Schmitz et al.,

2004; Grabowski and Kimbro, 2005) or for the evolution of body size, a trait

whose increase allows some degree of predator avoidance (Emmerson and

Raffaelli, 2004), while incurring metabolic costs (Kleiber, 1961).

Using a simple, Lotka–Volterra-based food chain model, Loeuille and

Loreau (2004) show how the evolution of plants, the evolution of herbi-

vores, or the co-evolution of the two, affect top-down controls. Their

results show, for instance, that evolution of increased plant defences under

nutrient enrichment weakens top-down controls. A more relevant question

from an agricultural perspective is whether insect pest evolution is affected

by nutrient enrichment and, if so, what are the possible consequences for

biological control. In Box 6.3, we extend the Loeuille & Loreau model

to answer this question, by showing that in a plant–herbivore–predator food

chain, evolution of herbivore defences decreases top-down controls exerted

on pests. One immediate implication for agricultural systems is that the

advantage gained on one side (increased yield through nutrient addition)

has a direct cost to an a priori unrelated ecosystem service (biological control).

In addition to these direct effects of nutrient enrichment on the selection

of traits, indirect evolutionary effects are also expected. Nutrient enrichment

affects the total number of species of the community (Chase and Leibold,

2002; Kassen et al., 2000) as well as the number of trophic levels

(Oksanen et al., 1981) so that interactions are globally changed within the

Implications for biological controlThe dynamics of food chains are fairly well understood. Particularly, when top-down controls dominate, nutrient enrichment increases the biomass of thetop compartment and of every alternate level below. The biomass of otherlevels remains unchanged or decreases (Hairston et al., 1960; Oksanen et al.,1981, see also Fig. 6.3, top). The strength of top-down controls strictly determine

100

20

0

40

60

80

100

120

150

P

Hevol

H

200Nutrient input

Bio

mas

s

250 300

Figure 6.3 Top, food chains: black corresponds to compartments whose biomassincreases, white to compartments not affected by the enrichment. For the ‘herbivoreevolution scenario’, enrichment selects larger values of the herbivore defence trait,depicted by an arrow on the right side of the herbivore level. Note that in the foodchain on the left, predators exert a biological control on the herbivores. Herbivoresdo not increase during enrichment. However, when evolution is possible, this is nolonger true. Enrichment allows the evolution of the herbivore to decrease thestrength of biological controls. Herbivore biomass then benefits from the enrich-ment and their biomass increases. Bottom, a simulation showing the herbivore bio-mass (in grey) and predator biomass (in black) for the ‘ecology alone’ scenario(dashed lines) and for the ‘herbivore evolution’ scenario (plain lines). Note that,starting from an initial nutrient input level of 100, the net effect of the evolutionof herbivores on the weakening of the biological control can be fully perceivedas the distance between the two grey lines. For the analytical demonstration and

Continued

parameter setting of Fig. 6.3, see Appendix C.

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BOX 6.3 Herbivore pest evolution under nutrient enrichment:Implications for biological control—cont'dthepossibilityofbiological controls, an importantecosystemservice (Costanzaet al.,1997). A crucial question is whether evolutionary dynamics strengthen orundermine top-down controls when nutrient enrichment takes place. Using anadaptive dynamics approach applied to a Lotka–Volterra food chain model,Loeuille and Loreau (2004) studied how the biomass variation pattern describedabove changed, depending on different plant–herbivore co-evolution scenarios.One of their conclusions was that plant evolution significantly weakens thetop-down control, because nutrient enrichment selects for higher defence levelsthatdecreaseherbivore impacts. This result is not especially relevantwithin the agri-cultural contextwhereplant trait isoftenmoreconstrainedbyhumanaction thanbydirect selection from herbivores, but a more interesting question arises about howtop-down control is changed considering the evolution of herbivore pests facingtheir natural predators. The action of such predators, that is, the biological control,may changedue toherbivore evolution, asmodifiedbynutrient additions.Westudythis issue by extending the model of Loeuille and Loreau (2004), adding a predatorlevel and studying theevolutionofherbivoredefences. In AppendixC,we showthatthe weakening of top-down controls observed in the original model also applies insuch longer food chains. That is, herbivores evolve higher anti-predator defencesduring the enrichment, which decreases the biological control (see Fig. 6.3). Theresult is fairly general. It does not depend on the trade-off functions that are used,just requires that thedefensivetraitof theherbivorehasanallocationcost. Identityofthe defensive trait can be of many types, from chemical anti-predator defences(development of toxicity) to changes in behaviour (increased vigilance, use ofrefugia, etc.).As inLoeuilleandLoreau (2004), co-evolutionscenarios—whether theyare plant–herbivore, herbivore–predator co-evolution, or co-evolution of the threespecies—are likely to yield much more complicated results. Note however thatstudyingherbivore evolution is a logical first stepgiven theagricultural context. Her-bivore pests are often insects or small invertebrates, whose evolution has beenshown to be very rapid in response to agricultural change (Gould, 1991; Carrièreet al., 2010): evolution of species at other levels of the food chain may very wellbe slower.

364 Nicolas Loeuille et al.

network. Because of such changes, the diffuse co-evolution of the entire

ecological network is likely to be affected. It is, of course, more difficult

to predict how such changes in the network will impact evolutionary

dynamics, but some recent models have linked diffuse co-evolutionary

dynamics to community structure or ecosystem functioning. Loeuille and

Loreau (2005) explain that, in a food web based on predator–prey body size

co-evolution that considers interference competition, nutrient enrichment

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365Eco-Evolutionary Dynamics in Agroecosystems

yields more trophic levels and diversity. Brannstrom et al. (2012), in a related

model, show that it also affects the strength of disruptive selection on body

size in ecological networks, thereby favouring ecological speciation. Results

of increased diversity, numbers of trophic levels and of changes in con-

nectance are also found in a co-evolutionary model that relies on many phe-

notypic traits evolving independently (Drossel et al., 2001). Note that the

optimistic message that may be inferred from these studies—that nutrient

enrichment increases community diversity—should be taken with caution,

because species niches in these models rely on a single energy axis, so that

enrichment simply inflates the available niche space for diversification. In

nature, at high nutrient loads, negative side-effects such as toxicity or anoxia

ultimately come into play to halt such increases in diversity (Harnik et al.,

2012; Justic et al., 1995).

3.2. Chemical warfare in agricultural landscapes:The ecological and evolutionary consequences ofpesticide use

A second disturbance that creates important selective pressures is the use of

insecticides, herbicides, or other chemical pesticides. While usually targeted

at specific pests, pesticide toxicities are usually tested on just a few other spe-

cies and they may have side-effects on a far larger number across different

trophic levels (Crowder et al., 2010; McMahon et al., 2012; Robinson

and Sutherland, 2002), some of which, such as pollinators, may provide

important ecosystem services (Klein et al., 2007). Direct effects can be

observed when sensitivity to the pesticide evolves due to selective pressures

exerted by chemicals. Alternatively, because this disturbance propagates

through the ecological network and affects its structure and diversity, it

modifies the selective pressures indirectly throughout the community, alter-

ing the course of diffuse co-evolution.

The use of pesticides is an important disturbance associated with agricul-

ture on local to global scales (Matson, 1997) and by generating extra mortality

on wild organisms, pesticides have detrimental consequences for ecological

dynamics and overall ecosystem functioning (Matson, 1997). Impacts have

been noted on many groups such as birds (Carson, 1962), pollinators (Potts

et al., 2010,) and natural enemies of pests (Bianchi et al., 2006; Geiger

et al., 2010). Decreases in such groups erode their associated ecosystem ser-

vices (Bianchi et al., 2006; Potts et al., 2010), yet few studies have considered

how associated evolutionary dynamics affect ecological networks.

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366 Nicolas Loeuille et al.

The potential for an evolutionary response to pesticides varies among wild

organisms associated with agricultural landscapes. Examples of rapid evolu-

tionary responses abound in the agricultural literature (Palumbi, 2001;

Thrall et al., 2011), where resistance to pesticides is a classical textbook exam-

ple of how continuous selection pressures on a trait with generally simple

genetic determinism results in rapid allele change. The international survey

of resistance in weeds (http://www.weedscience.com) listed 217 species

(129 dicots and 88 monocots) that had evolved resistance to 21 of the

25 known herbicide sites of action and to 148 different herbicides.

A similar pattern is observed in insect pests, with an extremely rapid increase

in the number of species showing resistance to at least one insecticide since the

1950s (Mallet, 1989). Approximately 8000 cases of resistance to 300 insecti-

cide compounds are now reported inmore than 500 insect species (Arthropod

Pesticide Resistance Database; http://www.pesticideresistance.com; Whalon

et al., 2008). Similarly, 300 cases of field resistance to 30 fungicides have been

reported in250 species of phytopathogenic fungi (FungicideResistanceAction

Committee database; http://www.frac.info; Bourguet et al., 2013). That evo-

lutionary responses are less well-known in non-pest organisms could simply

reflect the fact that they are largely ignored in such instances (Pelosi et al.,

2013), because of the lesser economic concern they represent.

It is, however, not expected that all species will evolve in response to

pesticides: even when a variable and heritable response trait exists, some pro-

cesses can limit its effective evolution. Plastic responses can play a role in the

emergence of resistance to pesticides, by alleviating fitness costs (Hendry

et al., 2011). For organisms with a plastic behaviour that can simply avoid

crops with insecticides selective pressures for increased resistance will be

weak. Such exploitation of alternative habitats can, however, incur indirect

costs, for instance when such habitats have fewer resources. When plasticity

actually impedes the evolution of pesticide resistance, it still raises important

evolutionary questions: namely, when do we expect the evolution of such

plasticity and what are the associated costs? A general idea is that plasticity is

favoured when selective constraints are variable in space and time (Agrawal,

2001; Lind and Johansson, 2007). Because pesticide use may be highly var-

iable in time, being most prevalent during infestation periods, and because

semi-natural habitats are not directly treated, creating a spatial heterogeneity,

plastic responses should commonly evolve. This view is supported by the

existence of many plastic behaviours allowing some resistance to pesti-

cides (Gould, 1991). Changes in the oviposition and host fidelity behaviour

of corn rootworm, for instance, allowed it to counter crop rotations

(Gray et al., 2009).

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367Eco-Evolutionary Dynamics in Agroecosystems

Demographic constraints, such as population density or generation time,

also affect the evolutionary response to pesticides. This idea can be illustrated

using evolutionary rescue models (Bell and Gonzalez, 2011; Gomulkiewicz

and Holt, 1995). A population facing a disturbance can go extinct when the

extra mortality decreases its per capita growth rate sufficiently. On the other

hand, selection of less vulnerable phenotypes will occur, increasing average

fitness. Evolutionary rescue is determined by the race between these two

processes (Gomulkiewicz and Holt, 1995): extinction or rescue depends

on the initial density of the population and on the generation time of the

species. Small populations or long-lived organisms, having small variability

in the response trait, are most at risk. Their densities following pesticide use

may fall under a critical value, where demographic stochasticity increases the

risk of extinction (Gomulkiewicz and Holt, 1995). However, species whose

effective population size is large (therefore possibly harbouring high genetic

variability), or generation times are short (so that new mutations can appear

more readily), may escape extinction when adaptation happens fast enough

to increase average fitness values before the population viability threshold is

reached. Most pests probably fall in this category, as they have short gener-

ation times and high population densities. Pollinator populations, on the

other hand, have lower densities and longer life cycles. Evolutionary rescue

therefore provides a potential explanation for the often-observed selection

of resistance to insecticides in pests (Bourguet et al., 2013; Mallet, 1989),

while many pollinator populations collapse (Beismeijer et al. 2006). Another

example is the evolution of pesticide resistance in fungi, which is determined

by the pathogen’s evolutionary potential, as set by its life cycle and repro-

duction system (Barrett et al., 2008; McDonald and Linde, 2002).

Most existing works related to pesticides tackle the conditions for the

evolution of higher resistance in pest species (Bourguet et al., 2013;

Gould, 1991; Mallet, 1989; Thrall et al., 2011), reflecting the huge eco-

nomic costs incurred (Palumbi, 2001). Evolutionary models can then sug-

gest managements that delay the emergence of these resistances, such as

high-dose/refuge or pyramid strategies that use the ensemble activity of

multiple compounds (Bourguet et al., 2013). A complete review of this

aspect is beyond the scope of the present chapter and has been covered else-

where (Bourguet et al., 2013; Carriere et al., 2010). These approaches, how-

ever, mostly focus on one species, the targeted pest, ignoring the broader

ecological network (Tabashnik et al., 2004; Vacher et al., 2003).

Many observations suggest that the community context does affect the evo-

lution of resistance. For example, resistance to insecticides in mosquitoes neg-

atively affects the resistance toparasites, so thatevolutionarydynamicsultimately

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368 Nicolas Loeuille et al.

depends on the overall parasite loadwithin the population (Agnewet al., 2004).

Resistance in thepeach-potatoaphid increases its sensitivity topredatorsdue toa

decreased response to alarm pheromones (Foster et al., 2003). Through such

ecological costs, the predation context affects evolution of resistance. Evolution

of resistances also affects pest densities, allowing them to attain large numbers,

and theseeffects canpropagate throughecologicalnetworks, creating top-down

or bottom-up effects. Because pests interact both directly and indirectly with

many other species, density-dependent effects will eventually impinge on the

eco-evolutionary dynamics of other wild species of the community.

We have little information on such eco-evolutionary dynamics within the

pesticide context, but another corpus of models, developed in fisheries sci-

ence, have studied the evolutionary dynamics of communities in which a sub-

set of species faces extra mortality, via harvesting. Even though these models

are not explicitly designed for agricultural systems, they are often stated in gen-

eral equation systems that can be applied outside the fisheries context,

although such an analysis is still restricted to the mortality effects of pesticides.

Other sub-lethal effects (e.g., modification of fecundity or of consumption

rates) require more explicit and detailed models.

Abrams and Matsuda (2005) studied a two-species model that we have

translated to a herbivore pest and its host plant, where the former suffers

extra mortality (e.g., due to pesticides) and plant nutrient uptake rate evo-

lves. Nutrient uptake rate is supposedly positively correlated with plant

vulnerability, because of an allocation trade-off between defences and

growth. The model shows that herbivore density can increase when its

own mortality increases. This counter-intuitive result occurs because herbi-

vore density initially decreases, causing selection of higher vulnerabilities of

the plant. Because the plant is more vulnerable, the density of her-

bivore subsequently increases. This latter positive evolutionary effect even-

tually exceeds the direct negative effects of extra mortality (Fig. 2 in Abrams

and Matsuda, 2005). This result has interesting agricultural implications

because it suggests an evolutionary mechanism through which pesticides

eventually increase pest abundances. Abrams and Matsuda suggest that

plastic responses could produce similar qualitative trends. The model would

also be suitable to describe the evolution of wild plants consumed by

the pest. When increasing pesticide application, pest density increases until

a threshold is reached, but beyond this value the pest abruptly goes extinct.

In a four-compartment model that can be translated as two plants sharing a

limiting nutrient and a pest herbivore, Abrams (2009) again studied how the

responseof thepest toextramortality is affectedbyadaptation in theplant. Plant

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369Eco-Evolutionary Dynamics in Agroecosystems

evolution can again create an increase in pest density when pesticide is used.

Compared with the simple food chain model of Abrams and Matsuda

(2005), the evolution of the second plant species extends the range of pesticide

use thepest can sufferbeforegoingextinct.Apossible agricultural implication is

that evolution of a lesser vulnerability in wild plants can improve the existence

conditions of a pest confrontedwith pesticides.While an exact test of such pat-

terns has not been done, existingdata comparing the evolutionofweedswithin

fields and out of fields can provide an interesting point of comparison (Ellstrand

et al., 2010). Evolutionary effects of extramortality can also affect the stability of

natural communities, whichmay increase or decrease, depending on howpes-

ticide impacts relate to population densities (Witting, 2002).

Another important result of fisheries research is that increasedmortality can

select for earliermaturationand smaller body sizeatmaturation (Allendorf et al.,

2008; Barot et al., 2004; Enberg et al., 2009; Law, 2000). Because of highmor-

tality, those who mature faster are more likely to reproduce and are therefore

favoured. Similarly, because the pesticide use exerts extra mortality, the same

selective mechanism may apply, with fast-reproducing strategies being

favoured, although we could not find any experiment or observation that

has tested this idea.Anobvious difficulty lies in the fact that confounding factors

exist: having a fast life cycle in the first placemakes itmore likely for a species to

be a pest, and it may also facilitate the evolution of resistance.

This prediction of selection for earlier reproduction may be dependent on

the community context. Gardmark et al. (2003) used a model in which a prey

populationconsistingof three-ageclasses suffers extramortality fromharvesting

and is consumed by a predator species. They studied the evolution of the age at

reproduction (age 2 or 3; age 1 being juveniles), depending on the stage being

predated. Substituting the prey species with a pest species and harvesting with

pesticide extra mortality, in the absence of predation evolution of early matu-

ration is favoured.However, reproduction is delayed to age 3when extramor-

tality acts on age 1 and predation on age 2. Thus, the two mortality-related

selective pressures can interact antagonistically (Gardmark et al., 2003).

Importantly, it is unlikely that only pests are affected by pesticide use, and

impacts on other members of the network are likely to modify associated

ecosystem services as a result. To illustrate this, we modelled a plant–

herbivore pest–predator food chain where both the pest and the predator

are vulnerable to pesticides. For the sake of simplicity, we considered that

only the herbivore evolves in response to pesticides and that evolution of

resistance incurs a cost in growth rate (allocation cost). The model and prin-

cipal results are detailed in Box 6.4.

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BOX 6.4 Evolution of resistance to insecticides in a tri-trophicfood chainWe model a three-species community consisting of a plant attacked by a herbi-vore pest, consumed by a predator and in which insecticide use affects both theherbivore and the predator. Detailed equations and analysis of the model aregiven in Appendix D. We assume that the herbivore evolves in response to insec-ticide use. Its susceptibility incurs an allocation cost: when susceptibilitydecreases, herbivore reproduction decreases (Carrière et al., 1994). A possiblemechanism is that conversion efficiency is reduced, due to an allocation in detox-ification metabolism (Després et al., 2007). We study three different communityscenarios: the predator is absent from the community, the predator has a low sus-ceptibility to insecticides, and the predator has a high susceptibility to insecti-cides. We use the adaptive dynamics framework to study the impact ofherbivore evolution on the three-species food chain (see Appendix D).

Figure 6.4 shows how evolved resistance changes when pesticide use(parameter l) increases. Figure 6.5 displays the variations in densities for the threecommunity scenarios and depending on whether herbivore evolution happensor not. As intuitively expected, resistance is selected for at higher inputsof insecticides (Fig. 6.4). Herbivore adaptation delays the extinction of both

Figure 6.4 Influence of insecticide intensity on the value of herbivore susceptibil-ity trait at the evolutionary equilibrium, for the three different community scenar-ios: ‘No predator’, ‘Low susceptibility of predator’ and ‘High susceptibility ofpredator’. Parameters values: r¼2; I¼0.2; w¼0.2; a¼0.5; dh¼0.5; g¼0.5;dp¼0.1; c(s)¼exp(0.1 s); j(s)¼exp(0.3 s); Low h¼0.1; High h¼1. For parameterdefinitions and details of the equations, see Appendix D.

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BOX 6.4 Evolution of resistance to insecticides in a tri-trophicfood chain—cont'd

the herbivore and the predator with increasing load of insecticides (Fig. 6.5).However, when the predator is not strongly affected by pesticides,lessevolved resistance appears in the herbivore. This last result can be explainedby density-dependent mechanisms: when the predator has a low susceptibil-ity, the plant is at a higher density because of the predator exerts top-downcontrol on the herbivore. It is then beneficial for the herbivore to be moresusceptible (but with a high conversion efficiency) as the efficient consump-tion of abundant plants offsets mortality costs. When the predator is highlysusceptible to pesticide, its density is decreased or it goes extinct, so that thedensity-dependent advantage disappears. Community context and evolution-ary dynamics thereby interact to determine evolution of pests subjected toincreased mortality.

Figure 6.5 Densities of the three species at the ecological equilibrium (left) or eco-evolutionary equilibrium (right) for different intensities of insecticide applicationand for three community scenarios. Continuous lines show plant density V0,long-dashed lines herbivore densityH0 and dotted lines predator density P0. In thesegraphs, all species coexist in areas labelled ‘1’, ‘2’ indicates plant–herbivore coexis-tence and ‘3’ parameter sets for which only the plant exists. Parameters values arethe same as in Fig. 6.4. When there is no evolution the susceptibility trait for theherbivore is fixed at 0 (alternative values do not alter the qualitative results).

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372 Nicolas Loeuille et al.

Overall, resistance levels increased with insecticide use, but predator

presence decreased the evolved level of resistance and pest density. There-

fore, conservation of predators not only maintains an important ecosystem

service through biological controls, but it may also limit the evolution of

resistance in the pest. Note, however, that this double service is lost when

the predator is highly sensitive to pesticides. Such effects are produced by the

interaction between density-dependent ecological processes (here, preda-

tion) and trait dynamics due to species evolution. They affect both the mag-

nitude of species responses and the qualitative direction of such responses, as

reflected in densities and trait values.

In more complex networks, evolutionary effects of pesticide can prop-

agate through many trait-dependent and density-dependent pathways

(Wootton, 1994; Yodzis, 2000). Predicting the associated diffuse evolution

is difficult, especially because most models are based on trophic interactions,

while pesticide use may also affect non-trophic compartments (e.g., pollina-

tors). The interplay between different interaction types is a growing research

area in ecology and evolution (Altermatt and Pearse, 2011; Fontaine et al.,

2011). Antagonistic and mutualistic interactions influence each other,

both through ecological, density-dependent effects and through selective

pressures (Adler et al., 2012; Herrera et al., 2002). This is particularly rele-

vant for agricultural landscapes, whose sustainability relies on ecosystem

services that come from different interaction types. Biological control is

linked to trophic interactions acting on pests, while pollination services rely

on mutualistic interactions. Pesticides may affect all different interaction

types, coupling the different ecosystems services through eco-evolutionary

dynamics.

3.3. Effects of altering species composition and relativeabundance of species in agricultural landscapes

A third disturbance corresponds to changes in habitat, species abundance,

distribution, and composition due to agricultural activities, which can have

multiple effects. By developing open landscapes at the expense of natural

habitats, agriculture modifies landscape composition; at more local scales,

community diversity and composition are also affected. Agricultural com-

munities usually contain just a few dominant species (the cultivated ones),

represented by a few artificially maintained genotypes (Purugganan and

Fuller, 2009), so the base of the food web changes differs markedly from

the natural state that typically contains many plant species.

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373Eco-Evolutionary Dynamics in Agroecosystems

The impact of crop management practices on communities is well-

known: for instance, changes in the relative abundance of crop and weed

plants can generate bottom-up effects changing arthropod community

structure (Lenardis et al., 2011). On the other hand, insects can directly

reduce weed biomass in favour of the crop biomass, and vice versa

(Maron and Crone, 2006; Norris and Kogan, 2005). Such changes in species

biomass at the different trophic levels alter selective pressures, which, in turn,

affect the degree of specialization in species interactions.

The topic of specialization has recently attracted considerable attention

from community and ecosystem ecologists (Forister et al., 2012; Poisot

et al., 2011) and at least three factors influence the relative fitness of gen-

eralists and specialists have been identified. First, evolution of specialization

depends on trade-off shapes, with weaker trade-offs, that is, low cost of

switching from one interaction to another, favouring generalists (Levins,

1962; Rueffler et al., 2006b; 2007). Second, specialization can arise from

ecological opportunities created in complex networks (Forister et al.,

2012), such as plant-mediated enemy-free space in tri-trophic systems

favouring specialization in herbivore insects (Singer and Stireman, 2005).

Third, temporal variation in fluctuating and highly disturbed environments

favours generalists while environmental constancy favours specialists

(Levins, 1968, Futuyma and Moreno, 1988). Existing models predict spe-

cialization under high resource predictability (little or no seasonality),

abundance and diversity (Roughgarden, 1972; Van Valen, 1965). Behav-

ioural selectivity and plasticity, however, modulate the effects of spatial and

temporal heterogeneity and can promote the evolution of specialization

(Poisot et al., 2011). In a mathematical model, Ravigne et al. (2009)

suggested that the joint evolution of local adaptation and habitat choice

leads to specialization regardless of the shape of trade-offs, provided that

the cost of habitat selection is not too high.

The general trend of agricultural intensification has been simplification

in terms of reduced diversity of species (few crop species, mostly one crop

plant per field), genes (few varieties, genetic homogeneity within each cul-

tivar and mostly one cultivar by field), and management practices (few active

ingredients or tillage systems). Consequently, the transition from traditional

to intensive agriculture should favour specialized interactions, although this

may be partially mitigated by crop rotation.While these ideas require exper-

imental investigation, two lines of observational evidence provide empirical

support. At the species level, outbreaks of pests feeding on the dominant

crop happen more readily in these landscapes. The apparently strong

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374 Nicolas Loeuille et al.

modularity (i.e., weakly interlinked subsets of species, strongly connected

internally) observed in species networks within agricultural systems suggests

a high degree of specialization (Macfadyen et al., 2011). Traditionally, mod-

ular structure has been thought a stabilizing factor, slowing down the spread

of disturbance between modules (May, 1972; Krause et al., 2003, but see

Allesina and Tang, 2012), but this has been challenged recently, as it appears

that although modularity can enhance stability in trophic networks whereas,

a highly connected and nested architecture, resulted from generalized inter-

actions, can promote community stability in mutualistic networks

(Bascompte et al., 2006, Thebault and Fontaine, 2010).

Specialization has important implications for biological control in agri-

cultural systems. Traditionally, it has been assumed (Howarth, 1991) that an

effective biological control should be highly host-specific, based on the

resource concentration hypothesis that consumers are more likely to prefer

the most abundant resource (Root, 1973). Experimental and theoretical evi-

dence, however, suggests that in the case of pest predators, generalists can

generate better pest control services (Flaherty, 1969), especially in pest-rich

communities where generalist predators maintain low prey populations via

resource partitioning (Symondson et al., 2002).

Less effort has been devoted to the study of biological weed control in crop

fields. On the one hand, the introduction of insects has been proposed as a

mechanism of weed control (McEvoy 2002), but it has also been hypothesized

that the top-down effects of a generalist insect on a crop can be diluted via the

presence of weed species (Root, 1973). Therefore, a generalist insect–weed–

crop system may be easy to manipulate to favour a crop or other beneficial

plant species. What remains unclear, however, is how eco-evolutionary

dynamics can affect specialization of the herbivore and hence, biological weed

control. A simple model of the interaction between a herbivore pest and two

plant species, one crop and one weed, competing for a common resource

(Box 6.5) can be proposed here, in line with the diamond shape commonly

applied in general ecological scenarios (Armstrong, 1979; Leibold, 1996). This

structure is also ideal for studying the interaction between weeds and insect

pests. Indeed, experimental observations suggest that insect pest outbreaks

may be negatively affected by the presence of weeds, because they may host

predators or parasitoids (Altieri, 1981). All else being equal, results of the dia-

mond model suggest that a crop species, despite being a poor resource com-

petitor, can coexist with a competitive weed, due to its herbivore-resistance

mechanism favoured by breeding efforts. Biological weed control in such a

system can be readily achieved via fertilization (see Box 6.5). The generated

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BOX 6.5 Degree of specialization of pests and its eco-evolutionaryeffects: Implications for weed and pest managementAmajor challenge inecology is tounderstandhowspeciescancoexistdespitebeinglimited by the same biotic or abiotic factors (Hutchinson, 1961; Tilman, 1982; Chaseand Leibold, 2003). In theory, species coexistence requires that the abundance ofeach species is limited by a different (biotic or abiotic) factor. A generalist herbivoremay represent this additional limiting factor and insure the coexistence of twoplantspecies competing for a common limiting resource (e.g., a mineral nutrient). Thisinteraction configuration matches fairly well the case of a generalist insect feedingon two plant species, a crop and aweed sharing a common limiting resource. Dam-age to the crop (weed) leads to decreased use of resources by the crop (weed).Resources not used by the crop (weed) theoretically become available to weeds(the crop), resulting in increased weed (crop) growth (Norris and Kogan, 2005).The question is how the eco-evolutionary dynamics of such a system can help todesign effective weed and pest management.

At low resource inputs (small I) the model predicts that competitively superiorweeds dominate because densities of herbivore would be insufficient to reverse theircompetitive advantage. At high resource input (high I), however, the crop dominatesbecause it can support ahighherbivoredensity that in its turn suppressesweed to lowbiomass. In other words, the pest functions as an agent of biological control on theweedathighfertilizationlevels.Therefore,anagronomistcanreadilyassureahighcropyield by merely increasing nutrient input and/or by inflicting a high mortality rate onthe weed (e.g., through herbicide use). The efficiency of this biological control mech-anism, however, can be compromised by the evolution of herbivore specialization.Astrongselectivepressureexists fora relative increase inherbivore’sspecializationrateonthemoreabundant/profitablecropspecies (seeAppendixE).Aconsequenceof thisevolutionofspecializationontheabundantcrop is that thepest-insectno longerexertsbiological control at high fertilization (Fig. 6.6B andC). Moreover, the evolution of spe-cialization enhances the top-down control on cropplants,which results in thepositiveresponse of the pest herbivore biomass to fertilization (Fig. 6.6A).

Nutrient input (I )

Cro

p bi

omas

s

Wee

d bi

omas

s

Pes

t bio

mas

s

Nutrient input (I )

0.0 0.2 0.4 0.6 0.8 1.0

0.0

A

B C

0.2 0.4 0.6 0.8 1.0

0.0 0.2 0.4 0.6 0.8 1.0

Figure 6.6 Illustration of changes in the biomasses of the pest, crop and weed spe-cies along a gradient of nutrient with (continuous lines) and without evolution (dot-ted lines). For the analytical demonstration and parameter setting, see Appendix E.

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376 Nicolas Loeuille et al.

increase in crop biomass, however, will likely affect the degree of specializa-

tion of the herbivore. The herbivore species responds adaptively to the relative

decrease of the weed population by increasing its specialization rate on the

most abundant crop population. Thus, the evolution of specialization, via

density-dependent mechanisms or plasticity, may ultimately compromise

the efficiency of biological control practices.

4. ACCOUNTING FOR SPATIAL HETEROGENEITIES:DISPERSAL, FRAGMENTATION, AND EVOLUTION

IN AGRICULTURAL LANDSCAPES

Take h

Sectio

Sectio

4.1

Sectio

4.2

4.1. Characteristics of agricultural landscapes, past,present and future

Large changes in agricultural landscapes have happened in the last 60 years:

on a global scale, cultivated areas (1.5 109 ha in 2003) increased by 13%

between 1961 and 2003 and irrigated land has doubled (Paillard et al.,

2010). In some regions of Asia and South-America, the ‘Green Revolu-

tion’ favoured the transition from subsistence agriculture to industrial agri-

culture, and similar changes took place in North America and parts of

Europe. Through increased inputs and breeding of modern varieties, the

productivity of the major crops improved, reducing malnutrition in some

ome messages and key references for section 4

nAgriculture and cropbreeding

Ecology and evolutionaryecology References

n Ecological intensification of

agriculture would lead to

more fragmented and

diverse agricultural areas.

Impact on fragmentation of

natural areas depends on the

land-sparing/land-sharing

choice

Habitat heterogeneity and

steepness of environmental

gradients affect eco-

evolutionary dynamics

Tscharntke

et al., 2005

n The spatial distribution of

fields affects gene flows from

cultivated species to wild

populations and among wild

populations. This

subsequently affects pest and

wild species evolution

The spatial organization of

the patches of a

metacommunity influences

gene flows and the

subsequent evolutionary

dynamics

Haygood

et al., 2003

Plantegenest

et al., 2007

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Take home messages and key references for section 4—cont'd

SectionAgriculture and cropbreeding

Ecology and evolutionaryecology References

Section

4.3

Conservation biological

control of pests aims at

promoting movement of

individuals from semi-

natural areas to fields.

Spillover of individuals from

crops to natural habitats also

happens. This impacts the

ability of species to adapt to

crop/natural habitats

Dispersal between patches

impacts community

composition and related

eco-evolutionary dynamics.

Amount of spillover and

habitat characteristics also

affect evolution of dispersal

and species adaptation to

new habitat

Rand et al.,

2006

Blitzer et al.,

2012

Section

4.4

Agriculture would benefit

from co-evolution scenarios

that would foster the

evolution of mutualistic

interactions between crops

and other organisms

Landscapes lead to a mosaic

of coldspots and hotspots of

co-evolution, linked by the

dispersal of either one or the

two species

Thompson,

1999

Hochberg

et al., 2000

Section

4.5

Agricultural activities create

gradients of productivity

due to fertilization

Eco-evolutionary dynamics

lead to higher diversity and

more complex networks

when gradients of

productivity are smooth and

in intermediate productivity

locations

Loeuille and

Leibold,

2008a

Doebeli and

Dieckmann,

2003

Section

4.6

Land sharing is likely to be

the safest option for

agriculture in terms of eco-

evolutionary dynamics

Based on metapopulation/

metacommunity models,

land sharing may better limit

the evolution of pest

resistance, maintain

mutualistic interactions,

maintain diversity

377Eco-Evolutionary Dynamics in Agroecosystems

regions and allowing other regions to reach food self-sufficiency (Evenson

and Gollin, 2003), albeit sometimes at the cost of increased social inequality

(Jarosz, 2012).

The industrialization of agriculture has had large impacts on the structure

(composition and spatial configuration, Turner, 1989) of the landscape, whose

composition (i.e., thenatureof landcovers/uses)hasbecome lessdiverse as local

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378 Nicolas Loeuille et al.

cropsor landraceshavebeenreplacedbya fewhighlyproductive cereals (wheat,

corn, and rice that now represent 40% of cultivated lands; Tilman, 1999). The

local intensificationof agriculture and increased inputs also reducedplantdiver-

sity in cultivated areas through improved weed control and reduced necessity

for set-aside fallows. The proportion of natural or semi-natural areas in land-

scapes were reduced and these areas became more fragmented, partly because

of an increase in cultivated area, but alsobecauseof land reallocationprocedures

to increase field sizes and facilitate mechanization. This has destroyed the edge

habitats that surrounded small fields and acted as corridors between different

parts of the natural habitat (Davies and Pullin, 2007), which, combined with

intensive agricultural practices in cultivated fields, sharpened the boundaries

between different habitat types.

The industrialization of agriculture has also altered the temporal dynam-

ics of land use, replacing stable perennial cultivated areas, such as orchards or

pastures, with arable land (Malezieux, 2011; Tscharntke et al., 2005) and a

fast habitat turnover through crop rotations. Asynchrony of sowing and har-

vest dates among crops further creates a ‘hidden heterogeneity’ of the crop

mosaic, such that a given landscape exhibits very different structures both

within and across years (Vasseur et al., 2013), including more frequent

periods with no vegetation cover (Malezieux, 2011).

The low sustainability of intensive agricultural development has been

widely documented, in particular based on its negative environmental impacts

(Flohre et al., 2011; Geiger et al., 2010). New paradigms have been proposed,

such as ‘sustainable agricultural intensification’ (Lee et al., 2006) or ‘ecological

intensification’ (Dore et al., 2011; Griffon, 2010; Malezieux, 2011). Based on

some principles of agroecology (Altieri, 1989, 1999), these approaches propose

that part of a response to global food security challenges, in the context of

increasing population size and climatic variability, relies on an increase of bio-

diversity at the agricultural landscape scale and within fields, and they should

also increase theprovisionof additional ecosystem services, beyond that of food

production (Macfadyen et al., 2012; Smith et al., 2012). In contrast to the lim-

ited rangeofmanagement techniquesused in intensive systems, at the landscape

level, this ecological intensification requiresmorediverse cropping systems (De

Schutter, 2010; Paillard et al., 2010). Within fields, higher levels of above-

ground diversity could be provided, either by growing spatial associations

of species, possibly in an agroforestry framework (Ratnadass et al., 2012;

Smith et al., 2012), or by increasing the diversity of crops over time by diver-

sifying rotations and/or including intercrops (Altieri, 1999). Increasing crop

intra-specific genetic diversity, by increasing thenumbers of genotypes per spe-

cies grown either within or across fields, is another suggested solution

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379Eco-Evolutionary Dynamics in Agroecosystems

(Goldringer et al., 2006;Tooker andFrank,2012).Because thedevelopmentof

ecological intensification involves large changes in agricultural management,

the associated (initial) economic costs must be considered. Recent studies sug-

gest that current trends in yield may not be sufficient to meet future demands

(Rayet al., 2013), so ideally ecological intensification shouldnotdecreaseyields

relative to the present situation (Tilman et al., 2011). To achieve this will

require considerable development in scientific knowledge, training of farmers

and more sophisticated landscape planning (Dore et al., 2011, Mediene et al.

2011). The predicted changes in agricultural systems are likely to increase

the diversity and the fragmentation of agricultural habitats, but whether the

converse will be true for natural and semi-natural habitats will depend on

the options taken during landscape planning.

Future agricultural management is often seen as a choice between land

sparing versus land sharing. With an increasing population to feed, the

land-sparing solution proposes to increase the local productivity of fields

by any means necessary, including high inputs, so that remaining ecosystems

can bepreserved. Following this approach, the resultwould be very intensive,

natureunfriendly, agricultural fieldsonone side, and sparedpatchesofwildlife

on the other side, the two being as clearly separated as possible (Fig. 6.7, left).

In the land-sharing solution, themanagement of agricultural ecosystems aims

at beingwildlife friendly (less intensive production, less inputs, etc.), with the

land being ‘shared’ by human and nature. Meeting increased food demands

then needs larger agricultural surfaces. If the land-sparing solution is chosen

(Fig. 6.7, left), the amplitude of environmental gradient increases, as all

Figure 6.7 Impact of land-sparing and land-sharingmanagement options on the spatialgradient of environmental conditions. On each panel, exploited areas are in dark, nat-ural systems in light. In the case of land sparing, the difference between the two habitatsis large and they are well separated in space. In the land-sharing scenario, agriculturalexploitation is made closer to natural conditions and agricultural fields are embedded ina matrix of natural habitats.

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380 Nicolas Loeuille et al.

patches, both natural and agricultural, are either extremely suitable or

extremely unsuitable for any species and the environment becomesmore spa-

tially aggregated. In the land-sharing situation (Fig. 6.7, right) these scenarios

are reversed. Agricultural fields are managed as lightly as possible so that var-

iations between the field and the surrounding ecosystems are minimized: the

ecotone is less extreme. Also, semi-natural habitat is kept within agricultural

landscapes, so that the environment is less spatially aggregated.

4.2. Consequences of agricultural landscape structurein terms of gene flow

By fuelling local genetic variability or, conversely, by introducing locally

maladapted genes, gene flows are major constraints on eco-evolutionary

dynamics. They interact with local selection processes to determine the

amount of evolutionary change. Also, they depend on the level of spatial

heterogeneity and species dispersal abilities. Here, we will give a brief review

of the ways the spatial organization can constrain gene flows, by considering

how gene flows have affected the domestication and artificial selection pro-

cess, how genes flow from cultivated species to surrounding ecosystems, and

how agricultural fragmentation affects gene flows among natural

populations. Finally, we will highlight the implications of gene flows for

the eco-evolutionary dynamics of agricultural pests.

Gene flows in cultivated–wild/weed complexes have long been recog-

nized as an important process for crop domestication and as a source of crop

improvement in traditional agriculture (de Wet and Harlan, 1975; Elias

et al., 2001). Most crops indeed have wild relatives with which they can

hybridize readily. Ellstrand (2003) lists 49 cultivated plant species for which

there is genetic evidence of spontaneous hybridization with wild relatives

among which wheat, corn and rice and 12 of the 13 most important crops

in the world hybridize with their wild relatives (Ellstrand et al., 1999).

Increases in cultivated areas and anthropogenic modifications of wild species

habitats have promoted the likelihood of new contact zones between crops

and related species (Crispo et al., 2011). When the crop and its relatives

co-occur, the structure of the agricultural landscape determines the amount

of crop-to-wild gene flow by affecting their relative spatial distributions. It is

only recently, however, with the use of GM crops, that the impact of the

spatial distribution of crops and their relatives on such gene flow has been

thoroughly investigated and that attempts have been made to quantify it.

Quantification has proved difficult because landscape-scale experiments

are logistically challenging and gene flow events are rare. Historical gene

flow estimates based on population genetic structure cannot be used

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381Eco-Evolutionary Dynamics in Agroecosystems

effectively because of the dynamical spatial distribution of crops over the

landscape (Holderegger et al., 2010; Sork et al., 1999). Most effort has thus

been devoted to measures of contemporaneous gene flow, as measured, for

example, in Brassica napus (Hall et al., 2000), Brassica rapa (Warwick et al.,

2003; Schafer et al., 2011), Beta vulgaris (Arnaud et al., 2010), and Agrostis

stolonifera (Snow, 2012). Experiments have also been conducted to estimate

the rate of crop-to-wild gene flow (e.g., for oilseed rape Jorgensen et al.,

2009), some of which have addressed the spatial distribution of plants

(Klinger et al., 1992; Massinga et al., 2003). Quantitatively, results depend

on the parental species and genotypes, especially their selfing rate, as well as

on the experimental design (e.g., plant densities and management) (e.g.,

Jorgensen et al., 2009). The amount of gene flow also depends on the rel-

ative abundance of the crop versus its wild relative (e.g., Giddings, 2000,

Lavigne et al., 2002). For species that are pollinated by wind or both insects

and wind, more insight concerning the impact of landscape structure may

come from models describing gene flows between fields planted with the

same crop (review in Beckie and Hall, 2008; Kuparinen, 2006). Qualita-

tively, most studies typically report leptokurtic gene flow curves, with most

cross-pollination events occurring over short distances and few events at

large distances. Modelling approaches indicate that gene flow will be max-

imized in landscapes where the wild relative populations are small and

nearby (Klein et al., 2006a,b). They also indicate that, in species with suf-

ficient long distance pollen dispersal (i.e., fat-tailed dispersal kernels, Kot

et al., 1996), gene flow with distant wild relative populations will largely

depend on the overall abundance of the crop over the landscape (Lavigne

et al., 2008), while for short distance dispersing species, it will mainly depend

on the proximity of the nearest field (Klein et al., 2006a,b). Larger relative

wild populations will also be less impacted because of the protection effect of

their local pollen cloud, that is, it is easier to find mates of the same species

(Lavigne et al., 2008; Rhymer and Simberloff, 1996).

Other landscape characteristics, such as its physical heterogeneity (barren

land, high hedgerows), may either promote or limit gene flow from crops

(Dupont et al., 2006; Manasse, 1992; Morris et al., 1994; Reboud, 2003).

Predicting in detail how the structure of the agricultural landscapes affects

crop-to-wild gene flow in insect-pollinated species is particularly challeng-

ing because pollinator movements may depend on the specific spatial distri-

bution of habitats (Hadley and Betts, 2012).

The fate of introgressed genes will depend on a balance between fre-

quency of gene introgression and selection for the ‘cultivated’ gene in the

hybrids and their progeny in the wild population (e.g., Nagylaki, 1975).

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382 Nicolas Loeuille et al.

If the introgressed gene is selected for, it can result in genetic assimilation and

possible losses of diversity in wild populations. Assimilation is also expected

for detrimental genes when migration rates are high (e.g., Felsenstein, 1976;

Haygood et al., 2003; Huxel, 1999; Nagylaki, 1975; Slatkin, 1985).

Introgressed genes can negatively affect the fitness of the hybrids or subse-

quent generations, placing wild populations at risk of extinction via demo-

graphic swamping (Haygood et al., 2003; Kwit et al., 2011; Levin et al.,

1996). Finally, if hybrids or their progeny exhibit higher fitness than their

parents, they may become weedy or invasive. Given the huge, and increas-

ing, areas that have been planted with cultivated crops and, given their abil-

ity to hybridize with wild relatives, invasions in natural areas are surprisingly

infrequently reported. Situations where hybridization led to invasive or

weedy species becoming agricultural pests are far more frequently reported,

possibly because more easily observed and of larger economic interest. These

include, for instance, Johnson grass, weedy rice and weedy beet (Ellstrand

et al., 2010).

Genetic assimilation is less easily observed because introgressed genes may

have little or no impact on the wild species phenotype (Rhymer and

Simberloff, 1996), yet reports of introgression of cultivated genes in the wild

species (or subspecies) are numerous (e.g., Arias and Rieseberg, 1994; de la

Cruz et al., 2005; Linder et al., 1998; Rieseberg et al., 1999; Warwick

et al., 2003; Whitton et al., 1997). The relatively infrequent reports of genetic

assimilation in wild relatives of crops or in invasive species that descend from

crops may be caused by the generally poor adaptation of crops outside fields

(deWet andHarlan, 1975) and the low fitness of hybrids with cultivated traits.

Indeed, low hybrid fitness, rather than promoting genetic assimilation of wild

populations, increases the risk of local extinction because of increased demo-

graphic stochasticity and maladaptation (Ellstrand, 1992; Simberloff, 1988).

The exact cause of extinction may then be difficult to determine a posteriori.

The modified spatial structure of the agricultural landscape also affects

existing gene flows between semi-natural habitats embedded within it, as

they are prone to shrinkage and fragmentation. In theory, this should isolate

wild populations from each other, increase their risk of extinction by a

simultaneous increase of demographic stochasticity and of genetic drift lim-

iting their evolutionary potential and favouring inbreeding depression (e.g.,

Simberloff, 1988; Stockwell et al., 2003; Young et al., 1996). These pro-

cesses are, in particular, expected in species that are specialists of semi-natural

environments, and which have low reproductive rates and limited dispersal

abilities (Ryall and Fahrig, 2006). For these species, habitat fragmentation

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383Eco-Evolutionary Dynamics in Agroecosystems

results in a loss of connectivity among populations (Kindlmann and Burel,

2008). The overall trend for reduced diversity in populations from fragmen-

ted landscapes has been confirmed using a meta-analysis, which further

highlighted that the level of diversity also decreased with age since fragmen-

tation (Aguilar et al., 2008). However, habitat fragmentation does not always

lead to genetic isolation of populations, and landscape genetics has helped to

clarify this issue (Manel et al., 2003). In some landscapes, semi-natural ele-

ments, such as hedges, can connect remnant forest patches, although the

effectiveness of such corridors largely depends on the species considered

(Davies and Pullin, 2007), and some species are also capable of movement

across the open habitats of agricultural landscapes (Kanuch et al., 2012).

In certain cases, gene flows by pollenmay even be enhanced in fragmented

landscapes, as is often the case for trees (Bacles and Jump, 2011), and in some

species with long distance dispersal, isolated individuals may receive more

diverse pollen than those growing in large groups (Ismail et al., 2012; Klein

et al., 2006a,b). A different pattern emerges for insect-pollinated species

(Hadley and Betts, 2012), for which the combined effects of habitat loss

and fragmentation tend to have negative impacts, probably because of pollen

limitation. Themost susceptible species are self-incompatible species that need

outcrossing from genetically distinct individuals (Aguilar et al., 2006).

Finally, some generalist species can live in both the semi-natural areas of

agroecosystems and in the fields, provided that pesticide intensity is not too

high (Rand et al., 2006; Samu and Szinetar, 2002) and for them habitat

opening through agricultural conversion is not necessarily negative and

may even make for more connected landscapes (e.g., Eriksson, 2012;

Neve et al., 2008).

While gene flows among natural populations affect their densities, and in

some cases threaten their conservation, gene flows are also modified for pest

species, with possibly large economic implications. Landscape structure

influences not only the dynamics of pests but also their evolutionary dynam-

ics (Burdon and Thrall, 2008; Plantegenest et al., 2007). Evolutionary con-

sequences of the spatial distribution of crops over the landscape have been

largely investigated in the case of the evolution of pest resistance to pesticides

(insecticides, fungicides, herbicides) and of variety resistances by pathogens.

The rapid increase of resistance in general is due to the widespread use of

pesticides that globally increases the selection pressure for resistance and

to the mode of action of compounds that focus on specific targets. Similarly,

varietal resistances by pathogens have broken down, due to the deployment

over large areas of a same monogenic resistance to a pathogen resulting in a

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384 Nicolas Loeuille et al.

rapid increase in virulence alleles: currently estimated at 4–6 years on average

in western countries (McDonald and Linde, 2002).

The spatial context also provides solutions for the management of the

evolution of resistance. Deployment strategies of plants or pesticides have

been developed to cope with the rarefaction of new pesticide molecules

and the high cost for producing varieties resistant to pathogens or, more

recently, insecticidal GM crops. When the pesticide resistance (or virulence

gene) is already present in the population, or when it appears quickly, as is

the case for many pesticides, spatial strategies aim to reduce the overall selec-

tion pressure and slow the progression of the resistance (respectively, viru-

lence) alleles or, even, to stabilize their frequency at a low level where

resistance (respectively, virulence) has an associated fitness cost (Bourguet

et al., 2013; Vacher et al., 2003). These strategies are based on establishing

mosaics (also named mixtures) of resistance genes/pesticide treatments over

the landscape at different spatial scales (within or between fields), so that pest

populations experience heterogeneous selection pressures. The optimal spa-

tial scale of patches within the mosaic depends on the dispersal ability of the

pest/pathogen (Bourguet et al., 2013). When the resistance is not already

present in the populations, as in GM crops, high-dose refuge strategies have

been proposed and tested (Bourguet et al., 2013; Gould, 1991; Lu et al.,

2012). High doses are meant to severely reduce pest densities to slow the

emergence of resistance, be it monogenic or polygenic, and to make resis-

tance functionally recessive. Refuges are most commonly areas that are not

treated with pesticides, not planted with GM crops, or not planted with vari-

eties resistant to a particular pathogen. They favour sensitive (respectively

avirulent) individuals of the pest species when there is a fitness cost to resis-

tance (respectively, virulence) (Bourguet et al., 2013; Carriere et al., 2010).

Dispersal of individuals between refugia and non-refugia zones is essential

(Carriere et al., 2010; Gould, 1998), so that their use actually alters gene

flows within the landscape.

Pests may also have hosts in natural populations, and wild or weedy

plant species that act as a ‘reservoir’ of inocula when the crop plant is absent

in the landscape are important for pathogen dynamics and evolution

(Burdon and Thrall, 2008). In fact, the optimal strategy to promote the

durability of varieties depends on the importance of reservoirs as sources

of inocula relative to field-to-field dispersal. When reservoirs are important

pest sources, high cropping ratios of resistant varieties should be promoted

because the pathogen population may be maintained at a low level (Fabre

et al., 2012). Wild or weedy plants may also include the host where pest

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385Eco-Evolutionary Dynamics in Agroecosystems

sexual reproduction occurs, which can accelerate pest evolution rates

(Plantegenest et al., 2007).

In practice, controlling the spatial structure of varieties or treatments over

landscapes can be difficult when different farmers own discrete parcels of land.

The scale of ecological processes then differs from the scale of management

(scale mismatch hypothesis, Pelosi et al., 2010). Some large-scale management

programmes have nevertheless been implemented in co-ordinated refuge

strategies for GM crops (Carriere et al., 2012) or in landscape designs aimed

at enabling efficient pest control (Bianchi et al., 2006).

4.3. Consequences of spatial modifications from ademographic point of view

Consequences of individual movements across agricultural landscapes go

beyond gene flows, and ‘spillover’ also modifies densities and species

interactions in both the cultivated and wild habitats. Most studies have

focused on movements from the wild to the cultivated habitats, with the

aim to protect the ‘integrity’ of crops (Rand et al., 2006): field edges, for

example, were long considered as a reservoir of weeds or diseases by

farmers (Norris and Kogan, 2000). Semi-natural habitats are also seen

as a source of pest enemies that can act as biological control agents.

The promotion of predator and parasitoid spillovers from semi-natural

areas into the crops is one aim of conservation biological control

programmes (Landis et al., 2000). Landscape management options

designed to promote biological control are based on increasing the prox-

imity of cultivated to semi-natural areas, so that pest enemies active in

fields may find additional food resources, refugia, or overwintering sites

in adjacent semi-natural areas (Bianchi et al., 2010; Brosi et al., 2008).

This increase of contact zones may, however, have evolutionary conse-

quences for populations inhabiting semi-natural areas, by changing their

interactions within their species assemblage (Knight et al., 2005).

A recent review also highlighted the frequency of the opposite direction

of spillover, from crops to semi-natural areas for a wide range of func-

tional groups (Blitzer et al., 2012), which will depend on species traits,

such as dispersal ability and attack rate on the pest prey (Chalak et al.,

2010), and on similarity between cultivated and semi-natural areas

(Opatovsky et al., 2010; Prevedello and Vieira, 2010). It may have dra-

matic consequences for ecological networks when crops are the domi-

nant land cover and when they harbour larger populations of organisms

than semi-natural habitats. Consequences are direct when herbivore

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386 Nicolas Loeuille et al.

spillover reduces the size of semi-natural plant populations (Louda et al.,

2003; McKone et al., 2001) or when generalist predators or parasitoids

considered beneficial in the fields attack non-pest herbivores (Gladbach

et al., 2011; Rand and Louda, 2006). Pathogens that build up large

populations in fields may also affect wild related species, and wild plants

may be affected by pollinator spillover from the crops. Indeed, cultivated

areas may provide rich resources for pollinators, as in the case of mass

flowering crops, such as bean or oilseed rape, and large pollinator

populations may build up (Hanley et al., 2011; Westphal et al.,

2003), which could benefit wild plant populations. However, such

changes of the pollinator community may also disfavour certain compet-

ing wild pollinator species and their associated plants in semi-natural

habitats next to crops (Diekotter et al., 2010). The interplay of such

changes in the composition of ecological communities and of gene flows

crucially affects the eco-evolutionary dynamics within agricultural land-

scapes, and is likely to affect species coexistence (Urban et al., 2008) and

the overall structure of interaction webs (Loeuille and Leibold, 2008a;

Rossberg et al., 2008).

Eco-evolutionary dynamics are also affected by changes in fragmenta-

tion due to agricultural activities. A directly impacted trait is dispersal: sev-

eral models show that the spatial heterogeneity of environmental

conditions (Mathias et al., 2001; Parvinen, 2002) or the fragmentation

of the landscape in patches of asymmetric size (Hanski and Mononen,

2011, Massol et al., 2011) favour the evolution of high variability in

intra-specific dispersal strategies. These theoretical results are consistent

with observations in natural populations (Hanski et al., 2004). Agricultural

activities, by modifying the level of fragmentation, change the selective

pressures that maintain such variation. Beyond a certain level of fragmen-

tation, positive evolutionary effects are likely to be compensated by

increased local extinctions, so fragmentation effects are likely to be nega-

tive overall. A recent model simulating the impact of climate change shows

that evolution of dispersal in a very fragmented landscape did not allow the

species to escape extinction (Kubisch et al., 2013), while it did at interme-

diate fragmentation levels. These examples suggest that fragmentation cau-

sed by agricultural activities not only has evolutionary—and sometimes

positive—implications for dispersal, but also should be seen in the light

of other disturbances.

The spatial heterogeneity created by agricultural activities raises the

question of the conditions under which a species that inhabits the natural

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387Eco-Evolutionary Dynamics in Agroecosystems

habitat will adapt to the agricultural field and be maintained there. Clearly,

this is of critical importance for developing sustainable agriculture and

accounting for species conservation. A suitable modelling framework for

addressing this question is the one developed for source–sink landscapes

(Pulliam, 1988), in which some areas can maintain populations even in

the absence of dispersal (sources), while others do not (sinks). The black-

hole sink model by Holt et al. (2003) is particularly relevant here: it shows,

in a fixed landscape, that if a population has a source (e.g., the natural habitat)

and a sink (e.g., surrounding agricultural fields), then, in the absence of hab-

itat selection, the population eventually adapts to survive in the sink. The

worse the conditions are in the sink, the longer it takes for adaptation to

occur, although higher dispersal from the source accelerates this rate, so

adaptation can be quite abrupt and rapid.

These results are particularly relevant for considering the erosion of

diversity in agricultural landscapes (Robinson and Sutherland, 2002).

From the black-hole sink model, one would expect that species from sur-

rounding habitats will eventually adapt to agricultural locations and that

diversity would be restored in the long run. Indeed, adaptations to new

agricultural landscapes have been observed in many species (Eriksson,

2012), such as those adapted to open habitats maintained by agricultural

activities. However, in high input, intensive agricultural fields, it is likely

that adaptation to the black-hole sink is delayed by the extreme conditions

in the sinks (i.e., the fields) so that diversity recovery is slower and may be

outpaced by extinctions (Gomulkiewicz and Holt, 1995). In the light of

the black-hole sink model, restoring diversity is much more feasible in tra-

ditional agriculture, where the conditions within fields and in surrounding

landscapes are not so different, so that adaptation to agricultural conditions

can happen faster. Restoration of diversity is important not only for the

conservation of species, as diverse networks of species interactions most

often have a better functioning and stability (Loreau et al., 2001) and

may in turn provide more efficient and stable ecosystem services to agri-

cultural fields (Crowder et al., 2010).

The black-hole sink model can also help to explain why cultivated spe-

cies so rarely invade natural ecosystems. Through artificial selection, culti-

vated species have often evolved a suite of traits that make them non-viable

in natural conditions (Meyer et al., 2012; Purugganan and Fuller, 2009), so

the surrounding ecosystems are strong black-hole sinks for them. Adaptation

and invasion of these habitats would therefore be very slow and, since the

evolution of these cultivated species is often constrained by human

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388 Nicolas Loeuille et al.

populations in ways that are unrelated to the sink conditions, there is no rea-

son to expect that the within-sink adaptation will ever take place.

4.4. Consequences of spatial structure for pairwiseco-evolution

Metapopulation models such as the black-hole sink model consider the evo-

lution of a single species in space, and by ignoring species interactions they

focus on spatial heterogeneities and fragmentation. However, most ecosystem

services on which agriculture relies, such as pollination or biological control,

depend on species interactions, so we need to know how the spatial effects of

agricultural activities affect pairwise co-evolution within communities.

A suitable framework for addressing this is the ‘geographic mosaic of

co-evolution’ (Thompson, 1999, 2005), which is based on the idea that

the conditions of co-evolution between two species are highly variable in

space. In some places, called ‘hotspots’, the two species interact strongly

and their evolution is mostly explained by their pairwise interaction. In

other places (‘coldspots’), the co-evolution is less tight due to local condi-

tions. The landscape consists of a mosaic of coldspots and hotspots, linked

by the dispersal of either one or the two species (Gomulkiewicz et al.,

2000). Several interaction types have been modelled in this framework, such

as host–parasite interactions (Nuismer and Kirkpatrick, 2003; Nuismer et al.,

2000), predator–prey interactions (Hochberg and van Baalen, 1998), or

mutualistic interactions (Gomulkiewicz et al., 2003).

A key strength of the geographic mosaic of co-evolution is that it goes

beyond this simple classification of interaction types, so mutualistic and antag-

onistic interactions need not be treated separately, and can be considered as

part of a continuum. The empirical situation that motivated the geographic

mosaic of co-evolution explains how this new approach arose (Thompson,

1999): in the interaction between a moth (Greya politella) and a plant of the

genius Lithophragma, the female moth transports pollen from flower to flower

when they oviposit within the corolla, thereby providing a basis for a mutu-

alistic interaction. The larvae subsequently eat a few flowers, decreasing plant

fitness. Therefore, the balance of the interaction, positive or negative, varies

depending on other local conditions. If many other pollinators are available,

the interaction may be seen as negative. If pollinators are rare, the interaction

can be seen as mutualistic, as the moth then provides a much-required service.

The nature, strength and degree of reciprocity of the pairwise co-evolution

will change accordingly, making the landscape a geographic mosaic of

co-evolution (Thompson, 1999). Theoretical works on the mosaic of

co-evolution have investigated this continuum of antagonistic and mutualistic

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389Eco-Evolutionary Dynamics in Agroecosystems

interactions, as well as their dynamics in space and time (Gomulkiewicz et al.,

2000, 2003; Nuismer et al., 1999, 2000, 2003).

This evolution along a continuum between antagonistic and mutualistic

interactions has potentially huge implications for agriculture. Sustainable agri-

culture would, ideally, rely on ecosystem services produced by mutualistic

interactions, so any co-evolution scenario that turns mutualistic interactions

into antagonistic ones would be detrimental. Considering this, the geographic

mosaic of co-evolution gives two important insights regarding the long-term

management of interactions. First, the long-term maintenance of mutualism

depends on local energetic constraints. Hochberg et al. (2000) investigated

how co-evolution makes the nature of a symbiosis change along gradients

of productivity. They showed that mutualistic interactions are expected in

poor environments, while richer places should harbour strong parasitism. If

one considers the interaction between plants and mycorrhizae or between

plants and soil microbes in general, the association should evolve to be mutu-

alistic in poor environments, but antagonistic when too much nutrient would

be added, due to evolution or adaptation of either plants or mycorrhizae. In a

geographic mosaic of co-evolution, nutrient addition could therefore indi-

rectly disrupt an important ecosystem service. Because the model by

Hochberg et al. (2000) is fairly general, the warning may seem speculative.

However, two models more tightly focused on this particular issue (De

Mazancourt and Schwartz, 2010; Thrall et al., 2007) and reviews of empirical

data (Verbruggen and Kiers, 2010) support this general idea.

The second important insight is that just a few strongly selecting hotspots

can drive co-evolution at the landscape scale (Gomulkiewicz et al., 2000),

and there is ample evidence of strong selection in agricultural systems

(Gould, 1991). If an agricultural field is a hotspot for the co-evolution of

an interaction, then co-evolution happening there would have important

ramifications for co-evolution within surrounding ecosystems. Returning

to the example of the maintenance of plant–microbe mutualism, this raises

a puzzling question: if fertilization within fields were to turn some mutual-

istic interactions into negative ones, might it also disrupt interactions in sur-

rounding ecosystems? An interesting example in which evolution under

agricultural practices may lead to large changes in wild populations concerns

pollination. In intensive landscapes, fragmentation of wild habitats and use of

insecticides have altered pollinator diversity and densities, disrupting the

ecological link between pollinators and wild plants. Under such conditions,

evolutionary models predict the selection of more self-fertilization in wild

plants (Massol and Cheptou, 2011), with large implications for the genetic

structure, further evolution and ecological dynamics of plant species.

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390 Nicolas Loeuille et al.

4.5. Beyond pairwise interactions: Consequences ofeco-evolutionary dynamics for community structureand composition

Agricultural development creates strong spatial heterogeneities in habitats

and species pools: semi-natural areas coexist with agricultural fields and

community structure changes markedly over space. Total diversity, the

number of trophic levels, and the overall productivity or biomass of the

community are all structural and functional aspects that vary between

the field, its margins, and the surrounding habitats.

Models accounting for spatial heterogeneities and variations in commu-

nity structure can help us understand their implications for the evolutionary

and ecological dynamics of species assemblages (e.g., community evolution

models along gradients of productivity may explain how spatial heterogene-

ities in nutrient enrichment affect community stability and complexity).

One relevant result of such models is that the emerging diversity eventually

maintainedthroughout the landscapedependsontheamplitudeof thegradientof

productivity. In a sympatric speciation model, Doebeli and Dieckmann (2003)

showed that maximum diversity is obtained at intermediate gradient amplitude.

If the initial gradient is shallow, increasing its steepness createsnicheopportunities

that are positive for the emergence of diversity. Increasing the steepness of the

gradient further leads toadjacent locationshaving increasinglycontrastedsettings.

Any dispersal then produces maladaptation, and such negative gene flows even-

tually impede the diversification process and select for generalists. This illustrates

that themanagement of productivity gradients, a central challenge in contempo-

rary agriculture, is likely to affect themaintenanceof species diversity at landscape

scales, as well as the scope for future speciation.

Such results donot seemtonecessarily dependon theprecise settingsof a par-

ticularmodel (seee.g.,Day,2001;Kawata,2002;Massol,2012): inaverydifferent

model based on predator–prey co-evolution, Hochberg and van Baalen (1998)

also showedhowtheproductivitygradientaffects themaintenanceofphenotypic

diversity.Thetotaltraitdiversitymaintaineddependedontheslopeoftheproduc-

tivity gradient and peaked in areas of intermediate productivity. In an unrelated

model, Loeuille and Leibold (2008a) studied the evolution of plant defences in

response to specialist and generalist herbivores, along a gradient of productivity.

As with the other models, regional diversity was higher when the steepness of

the productivity gradient was intermediate. Local diversity was again highest

and trophic structure most complex in areas of intermediate productivity. In

Box 6.6, we discuss more thoroughly how this model can be used to explore

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BOX 6.6 How enrichment and fragmentation interact, illustrationusing a plant defence-based evolutionary modelContemporary ecosystems are often exposed tomultiple stressors, but how theseinteract is usually difficult to assess. In the case of nutrient enrichment and frag-mentation, an important point is that local selective pressures are modified, butdispersal is also altered. Dispersal and local selective pressures interact in com-plex ways (Urban et al., 2008). Part of this interaction stems from gene flows thatconstrain the genetic variation on which natural selection can apply. Other mech-anisms are also involved, such as the modification of community structure due tothe incoming dispersal of new species or when dispersal creates density-dependent effects.

A key condition to tackle this interaction is to consider simultaneously com-munity structure, evolution and space. One example model that fits theserequirements is the one by Loeuille and Leibold (2008a), which studies the evo-lution of two types of plant defences. One defence is efficient against all herbi-vores, but has an allocation cost that decreases plant growth rate. The seconddefence type is efficient against generalist herbivores, but may attract specialistherbivores, therefore having an ecological cost. Space is divided in 12 patchesalong a gradient of productivity, connected by passive dispersal of herbivoresand plants. Evolutionary dynamics can yield diversification of the defensive strat-egies, but such a diversification critically depends on the level of dispersal. If dis-persal is very high, the plant population has only highly defended morphs (nodiversification), else three functional groups coexist within the metacommunity(low defences, intermediate defences, high defences). In the latter case, highlydefended morphs are in richer patches, lowly defended morphs in poor patchesand intermediate defences are observed mainly in intermediate patches. Thedegree of spatial overlap of these three functional groups depends again onthe level of dispersal in the metacommunity. As shown in Fig. 6.8, these evolu-tionary dynamics constrain the architecture of the local food webs.

Using these results (Fig. 6.8), it is possible to illustrate how fragmentation—implemented as a decrease of dispersal—and nutrient enrichment interact toaffect the local structure of food webs. Arrow 1 shows that the effects of fragmen-tation can be impressive while you stand in a location of poor productivity. Atfirst, fragmentation has a positive effect on diversity, because it increases spatialstructure along the productivity gradient. Subsequently, diversity is lost as onlylocally adapted morphs are maintained in the patch. Arrows 1 and 2 show howthis effect of fragmentation depends on the productivity of the patch: in a richpatch (arrow 2), fragmentation does not affect the local food web at all as suchpatches are dominated by highly defended morphs regardless of the dispersallevel.

Continued

391Eco-Evolutionary Dynamics in Agroecosystems

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BOX 6.6 How enrichment and fragmentation interact, illustrationusing a plant defence-based evolutionary model—cont'd

Similarly, effects of enrichment will depend on the level of fragmentation inthe system. If the dispersal is very small (arrow 3), increasing the richness of thepatch first increases diversity, but then decreases the complexity of the food web.Note that such a bell shape relationship between complexity and productivity iscommonly observed in ecology, in models as in experiments or empirical works(Kassen et al., 2000; Mittelbach et al., 2001; Chase and Leibold, 2002). However, itis not observed when enrichment happens in a very well-connected met-acommunity (arrow 4). Then, neither local diversity nor local trophic structureis modified by nutrient enrichment.

Figure 6.8 Foodwebarchitectureas constrainedby theevolutionofplantdefences inametacommunity. Eachweb is supportedby a limiting inorganic nutrient (triangle) onwhich plants (circles) feed, supporting herbivores (squares). The white circle is a plantspecies whose evolution is not considered. For the other (evolving) plant, the shade ofgreygives thequantityofdefence theplanthasatevolutionaryequilibrium, from lowlydefended (pale grey) to highly defended (black). At the start of the simulations, thewhite herbivore is a generalist feeding on both plants while the black herbivore spe-cializesontheevolvingplant species. The initial foodwebarchitecture therefore resem-bles the middle web of the left column. Arrows (1) and (2) show two contrastedfragmentationscenarios, arrows (3) and (4) showtwoenrichment scenarios, asdetailedin the text. Adapted from Loeuille and Leibold (2008a).

392 Nicolas Loeuille et al.

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393Eco-Evolutionary Dynamics in Agroecosystems

themanagement of existing interactions between nutrient enrichment and frag-

mentation, two common side-effects of agricultural activities.

4.6. Land sparing versus land sharing, from an evolutionarypoint of view

The pros and cons of land sparing versus land sharing have been discussed

in detail frommany different viewpoints: economic, ecological, agricultural,

political, etc. (Fischer et al., 2011; Green et al., 2005; Hodgson et al., 2010;

Tscharntke et al., 2005). Here, we ask whether one of these two solutions is

overall likely to perform better than the other, considering eco-evolutionary

dynamics. This is particularly important because these management strate-

gies involve large-scale modifications of landscapes over several decades, a

timescale over which evolution needs to be considered, given the strength

of selection agriculture imposes on natural systems. Before we explore the

benefits and costs of the two options, let us reconsider the spatial conse-

quences of the two choices, which differ in the amplitude of the environ-

mental gradient (large for the sparing strategy, lower for the sharing strategy),

and the level of spatial aggregation (high for sparing, lower for sharing)

(Fig. 6.7). Considering these characteristics, land sharing appears to be a

more sensible option in the light of eco-evolutionary models, for three main

reasons.

The first concerns the management of pest evolution. In the land-sparing

solution (Fig. 6.7, left panel), high pesticide inputs equate to high selective

pressures on the enemies of the cultivated species, so they are likely to evolve

resistance in just a few generations (Carriere et al., 2010; Gassmann et al.,

2009; Gould, 1991). The use of no-pesticide refuges is already widely

applied to fight the spread of resistance (Carriere et al., 2010), based onmany

evolutionary models (Bourguet et al., 2013; Tyutyunov et al., 2008; Vacher

et al., 2003). Under land-sharing management, patches of natural habitats

readily provide many refugia in which resistance will be counter-selected.

In the land-sparing choice, agricultural fields are aggregated, so high selec-

tive pressures and limited availability of and connectivity with refugia make

the appearance and maintenance of resistance much more likely.

The lessons of evolutionary models for the maintenance of diversity and

the complexity of communities also suggest that land sharing is a safer

choice. As evolution along a gradient of productivity of intermediate

steepness is more favourable to the emergence and maintenance of species

diversity (Doebeli andDieckmann, 2003), locations of intermediate produc-

tivities should contain maximum phenotypic diversity (Hochberg and van

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394 Nicolas Loeuille et al.

Baalen, 1998) and the most complex food webs (Loeuille and Leibold,

2008a, Box 6.6). The gradient associated with land sparing is likely to be

too extreme to maintain diversity, and intermediate locations are absent

(Fig. 6.7). Land sharing, on the other hand, provides a shallower gradient

with many locations of intermediate productivity, creating a spatial structure

that may harbour high diversity.

A third reason why land sharing may be a better solution is linked to the

results of the geographic mosaic of co-evolution. Due to gene flows, hot-

spots can drive the co-evolution at larger scales, even when rare, if selection

is sufficiently strong (Gomulkiewicz et al., 2000). Under land-sparing man-

agement, it is likely that this condition is met because very intensive agricul-

ture concentrates on part of the landscape, so selective pressures there will be

high for any species present. Hence, if these exploited locations are hotspots

for some species co-evolution—not an unlikely scenario because few species

will coexist in such high-input, fragmented landscapes, so their reciprocal

interactions will be relatively more important—these will influence

co-evolutionary dynamics throughout the landscape. Essentially, this means

that what is considered ‘land sparing’ may not really be sparing species

co-evolution in space.

Wemust point out, however, that the black-hole sink model (Holt et al.,

2003) gives a more nuanced answer concerning the relative benefits of land

sharing and land sparing. For instance, with the spatial heterogeneities

depicted on Fig. 6.7 for a species that is best suited to the natural area in

the land-sparing situation, the cultivated area is likely to be very unsuitable,

creating a very strong black-hole sink. Adaptation to the sink will be very

slow to arise, a drawback from a conservation point of view (although from

a land-sparing perspective, these areas were not designed to support biodi-

versity in the first place). In land-sharing scenarios, cultivated areas are more

wildlife-friendly and closer to natural conditions, so they may still be sinks,

but less strong. Adaptation should therefore be faster, allowing some species

to persist in the exploited areas. Now consider a cultivated species, for which

the reverse reasoning can be made. In land-sparing scenarios, this species will

never adapt to the natural environment, because it will be a very black-hole

sink, and because the species evolves through artificial selection. But in the

land-sharing scenarios, evolution of the cultivated species will have to

account for conditions closer to the ones existing in semi-natural areas.

Environmental conditions vary less spatially, so some of the cultivated spe-

cies may adapt to the natural locations or gene introgression may become

more frequent. Therefore, from an evolutionary point of view, it is quite

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395Eco-Evolutionary Dynamics in Agroecosystems

possible that land sharing eventually produces new ‘invasive species’ (or at

least large introgressions) in the embedded semi-natural areas.

5. PERSPECTIVES AND CHALLENGES

In this chapter, we have taken recent insights gained from general eco-

evolutionary models and applied them to agricultural perspectives. On the

one hand, attempting such an exercise represents a stretch because commu-

nity evolution models are general, while agricultural questions are precise

and embedded in real field situations. However, as exemplified by the

land-sharing versus land-sparing debate, there is a need for agriculture to

set a general strategy for sustainability to complement smaller-scale political

decisions. General questions and debates do also exist regarding agriculture

sustainability, how it is constrained, and community evolution models can

be useful when these general questions arise. It is therefore desirable to try to

adapt certain features of these mechanistic models to the agricultural context,

to obtain more appropriate answers to these questions.

The spatial heterogeneity of environmental conditions in agricultural

landscapes is often quite different from that assumed in community evolu-

tion models. Some locations are agricultural fields and some others are nat-

ural habitats and transitions between the two are quite abrupt. Variations in

location type and environmental characteristics should be modelled accord-

ingly, rather than on a continuous environmental gradient, a choice that is

often favoured in community evolution models (Dieckmann and Doebeli,

1999; Doebeli and Dieckmann, 2003; Norberg et al., 2012). Also, simply

varying environmental conditions may not be sufficient. In agricultural

fields, the cultivated species often account for most of the local biomass,

so to model efficiently such situations, it is important to have the dominant

species and species composition varying—again abruptly—across the land-

scape. The metacommunity framework is a suitable starting point for devel-

oping such evolutionary approaches (Urban et al., 2008), but adapting it

fully to suit agricultural settings more effectively will involve some further

efforts (Massol and Petit, 2013, Chapter 5 of this volume).

Another possible issue concerns the modelling of evolutionary dynamics.

Individual-based selection drives evolution in wild species, while cultivated

species are most often selected based on yield, a group property. In Box 6.2,

we study co-evolution, combining individual selection for the wild species

and group selection for the cultivated species in a simple two-species model.

The results differed greatly from the initial individual selection model, and

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396 Nicolas Loeuille et al.

the robustness of the simple two-species community decreased markedly.

To understand the co-evolution between cultivated species, based on

human choice, and wild species, these differences in selection regimes need

to be built into appropriate models in the future. If we can simultaneously

adapt landscape structure and selection regime to agricultural constraints, the

geographic mosaic of co-evolution could provide important new perspec-

tives on pest management (Bousset and Chevre, 2013).

Another necessary improvement is to gain better assessments of evolu-

tionary speed driven by artificial selection for cultivated species relative to

other species in the landscape. Superficially, it would seem that artificial

selection, with breeding programmes and genetic tools, would produce

faster evolution for cultivated species, with the evolutionary dynamics of

wild species lagging behind, but this is not necessarily true. In fact, the

development of new traits, new seeds or new species for agriculture is faced

with many constraints, ranging from the research time needed to develop

these, to the health regulations to sell them. Once on the market, it is in

the interest of companies that sell them to maintain them for a sufficient

amount of time and in a sufficiently large number of places, so that the initial

investment is compensated. Clearly, such marketing constraints favour the

replacement of many genotypes by just a few recently engineered (Bonneuil

et al., 2012). This decrease in genetic diversity in turns diminishes the poten-

tial for fast evolution in the future. Sociological issues may also be involved:

the reluctance of some populations regarding some engineering tools (e.g.,

GMO) has clearly slowed down their use in some parts of the world.

Similarly, legislation (e.g., on the exchange of seeds) also constrains the

speed of evolution for cultivated species. For all these reasons, evolutionary

speed may not be so fast as is commonly perceived for most cultivated

species. It is, however, abrupt, as it may introduce new genes or traits on

very large scales in very short times when new varieties are chosen. That

intensive agriculture still uses so much input emphasizes that artificial

selection has failed to provide cultivated species adapted to local conditions

and their variations. Evolution of wild species on the other hand is very var-

iable in speed and depends on the genetic variability, population size, mating

system and generation time of the organism involved (Frankham, 1996;

Leimu et al., 2006; Loeuille, 2010b; Soule, 1976). It is important to

bear in mind here that agricultural pests, whether they are soil

pathogens (Burdon and Thrall, 2009), insect herbivores (Carriere et al.,

2010) or weeds (Gould, 1991), all have large population sizes and short

generation time, allowing them to adapt rapidly to changes imposed by

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397Eco-Evolutionary Dynamics in Agroecosystems

agriculture. Many other species (pollinators, predators, etc.) do not have

such traits and therefore cannot evolve at a similar pace, which might help

to partially explain their declines in agricultural landscapes. The assessment

of evolutionary speed is clearly critical for the associated modelling and if

species are expected to evolve on similar time scales, a co-evolutionary

model is most appropriate. On the contrary, if a subset of species evolves

sufficiently slowly, it is possible to ignore their evolution for short or inter-

mediate term questions, decreasing the complexity of the model and increas-

ing its robustness.

The development of adequate models can help to address important chal-

lenges. One of them is the consideration of agricultural change in the context

of other global scale disturbances. A large literature exists on the effects of cli-

mate change, and the understanding of multispecies system responses to global

change has been the subject of considerable research (Hagen et al., 2012;

Ledger et al., 2013; O’Gorman et al., 2012; Raffaelli and White, 2013).

Understanding of the evolutionary aspects and dispersal constraints,

of climate change, are also just starting to emerge (Le Galliard et al., 2012;

Norberg et al., 2012; Kubisch et al., 2013;Moya-Larano et al., 2012). Because

agriculture drives changes in habitats, in species composition and in fragmen-

tation, it interacts strongly with climate change disturbances and joint studies

of these two aspects are clearly urgently needed. This interaction between

agricultural modifications and CO2-related disturbances also involves fertili-

zation management, and several experiments have shown how CO2 and

nutrient loads interact to create unexpected responses in wild communities

(Langley and Megonigal, 2010; Reich, 2009).

Finally, a lot of work remains to be done to understand how eco-

evolutionary dynamics affect our perception of ecosystem services. As

shown in Boxes 6.3 and 6.5, it is quite possible that the selective pressures

we exert on a community (e.g., nutrient enrichment or weed removal) affect

an a priori unrelated ecosystem service (biological control). Ecosystem ser-

vices constrain the future sustainability of agriculture, but also the extent

to which we can rely on ecological intensification (Dore et al., 2011;

Malezieux, 2011) or the benefits we would gain from land sharing.

Trade-offs existing between different phenotypic traits may naturally couple

different ecosystem services together: for instance, plant chemistry simulta-

neously affects the repulsion of some herbivores, the attraction of others

(Muller-Scharer et al., 2004; Van Zandt and Agrawal, 2004), the attraction

of enemies (Poelman et al., 2008), and of pollinators (Adler et al. 2006; Adler

et al., 2012). These chemical traits clearly couple pest management with

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398 Nicolas Loeuille et al.

pollination and biological control ecosystem services, and some defences,

such as tannins, simultaneously decrease plant growth rate and change the

litter degradability (Grime et al., 1996; Whitham et al., 2003). Evolution

then affects two ecosystem services: food provision, and long-term mainte-

nance of soil organic matter. In the light of evolutionary trade-offs, it is

important to develop models that investigate this coupling of various eco-

system services. Adequate community evolution models could be a great

step forward, and agricultural issues could therefore provide important inspi-

ration for the development of future theoretical models in evolutionary

ecology.

The implications of eco-evolutionary dynamics are increasingly being

considered in theoretical and applied ecology: The American Naturalist

recently devoting a complete issue on this subject, for instance. It opens

many new doors to important contemporary ideas whose social and eco-

nomic implications could be very valuable in the near future. For instance,

the development of adaptive dynamics models, and more generally the con-

sideration of eco-evolutionary feedbacks opened many new doors in our

understanding of anti-biotic resistances (Day, 2001), of the evolution of par-

asite virulence (Alizon and Lion, 2011) or of vaccination strategies (Gandon

and Day, 2007) in the field of biomedical research into human health. Sim-

ilarly, the development of eco-evolutionary approaches in an agricultural

context can unveil many new of the currently unknown consequences of

artificial selection, the management of crop pests, and the sustainability of

ecosystem services.

ACKNOWLEDGEMENTSIn addition to INRA, IRD, UPMC and CNRS that routinely support our research, the

authors acknowledge the financial support of the Region Ile de France (DimAstrea grant

allocated to N. L. and G. K.) and of the INRA metaprogram SMaCH (C. L.). We thank

Francois Massol and an anonymous referee for their interesting comments on an earlier

version of this text.

APPENDIX A. EVOLUTION OF THE INVESTMENTINTO NUTRIENT UPTAKE, EFFECTS ON

EMERGENT FUNCTIONING

The simple three-compartment model describes the recycling of a

limiting nutrient between primary producers (V), dead organic matter

(D), and the stock of inorganic nutrient (N). Details can be found in the orig-

inal article (Boudsocq et al., 2011). The nutrient is recycled within the

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399Eco-Evolutionary Dynamics in Agroecosystems

ecosystem via internal recycling rates: the primary producer mortality rate

(dv), the mineralization rate of dead organic matter (md), and the uptake

of the mineral nutrient by V (g). Nutrient inputs are fixed and occur in

organic (Rd) and mineral (I) forms. Nutrients are lost from the ecosystem

with fixed rates, respectively lv, ld, and ln for the V, D and N compartments.

lv accounts for losses of nutrients through fires in terrestrial ecosystems or

harvest in agricultural systems. Dead organic matter and its content in nutri-

ent can be lost through fires, erosion, and leaching (dissolved organic matter)

(ld). Mineral nutrients are lost through leaching or denitrification (ln).

Finally, if the limiting nutrient is nitrogen and if there are nitrogen-fixing

primary producers (leguminous plants, some cyanobacteria), we assume that

fixation leads to an inflow of nutrient proportional to the primary producer

biomass (fv). In addition, we define primary productivity as:

’¼ gNV þ fvV ðA1ÞThe system of differential equations reads:

dN

dt¼mdDþ I � gV þ lnð ÞN

dV

dt¼ gNV � dvþ lv� fvð ÞV

dD

dt¼ dvV þRd� mdþ ldð ÞD

ðA2Þ

Using Eq. (A2), the formulas for the non-trivial ecological equilibrium

can be derived:

N 0¼ dvþ lv� fv

g

V 0¼Rd

md

mdþ ldþ I� lnN

0

dvþ lv� fvð Þ 1�bð Þ

D0¼Rdþ I � lnN

0ð Þ dv

dvþ lv� fv

mdþ ldð Þ 1�bð Þ

ðA3Þ

with b¼ dv

dvþ lv� fv

md

mdþ ldthat can be considered as the recycling efficiency

of the ecosystem.

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400 Nicolas Loeuille et al.

We assume that the evolutionary dynamics is constrained by a trade-off

between the uptake of nutrient and nutrient losses by plants:

g¼ g0ebs

dv ¼ dv0ecs

ðA4Þ

s is the evolving trait that affects positively the investment into nutrient

uptake. b and c determine, respectively, the benefit and the cost of additional

investment into mineral nutrient uptake.

To predict evolutionary dynamics, we use the Adaptive Dynamics

framework (Dieckmann and Law, 1996; Geritz et al., 1998). In adaptive

dynamics, the relative fitness of a rare mutant sm in a population s at its eco-

logical equilibrium emerge from the demographic dynamics:

WVmsm,sð Þ¼ 1

Vm

dVm

dt

����Vm!0

¼ g smð ÞN0 sð Þ� dv smð Þþ lv� fvð Þ ðA5Þ

We thus get:

WVmsm,sð Þ¼ g0e

bsmdv0e

csþ lv� fv

g0ebs� dv0e

csm þ lv� fvð Þ ðA6Þ

To reach an evolutionary equilibrium the selection gradient must be

null:

@WVmsm,sð Þ

@sm

� �sm!s

¼ dv0ecs b� cð Þþ b lv� fvð Þ¼ 0 ðA7Þ

If lv> fv and c>b, there is thus a unique evolutionary equilibrium or

singular strategy:

s� ¼ 1

cln

b lv� fvð Þc� bð Þdv0

� �ðA8Þ

We focus here on this particular case (realizedR�) but two other cases arepossible: tragic R� (b> c, runaway evolution that leads to the asymptotic

extinction of the species) or explosive R� (c>b and fv> lv, the evolutionary

dynamics lead to a situation where the limiting nutrient accumulates up to a

level where it is no longer limiting).

Once the evolutionary singularity is determined (Eq. A8), it is possible to

determine whether this singularity is convergent (in the sense that selective

pressures favour strategies that are closer to the evolutionary equilibrium)

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401Eco-Evolutionary Dynamics in Agroecosystems

and whether it is invasible (in the sense that alternative strategies can invade

the evolutionary singularity). Convergence and invasibility in adaptive

dynamics emerge from the second derivatives of the relative fitness

(Geritz et al., 1998). Given that:

@2WVm

@s2

!sm!s!s�

¼ bc lv� fvð Þ

@2WVm

@s2m

!sm!s!s�

¼�bc lv� fvð ÞðA9Þ

it is shown that the singular strategy is convergent and non-invasible: it is a

continuously stable strategy (CSS). This means that the evolutionary dynam-

ics eventually lead to this strategy and stop there.

Finally, it is possible to show that the sign ofdN 0

dt¼ @N 0

@s

ds

dtis always neg-

ative (see appendix 4 in Boudsocq et al., 2011), so that the singular strategy s�

minimizes the availability of mineral nutrient and nutrient losses. Expressing

@’0

@s

� �s!s

� and@V 0

@s

� �s!s

� , it can also be shown that the CSS neither maximizes

primary productivity nor plant biomass (see Appendix 5 in Boudsocq

et al., 2011).

Finally, for the numerical simulations (Fig. 6.1A of the present chapter),we

use fv¼0.01 year�1, lv¼0.1 year�1, ln¼0.05 year�1, ld¼0.0077 year�1,

dV0¼0.275 year�1,g0¼0.0137 ha kg�1 year�1, md¼0.0766 year�1,

I¼6 kg ha�1 year�1, Rd¼6.7 kg ha�1 year�1, b¼1, c¼1.735.

APPENDIX B. MIXING GROUP SELECTION ANDINDIVIDUAL SELECTION IN

CO-EVOLUTIONARY MODELS

The model described in Loeuille et al. (2002) considers a nutrient

explicit version, including recycling processes, of plant–herbivore

co-evolution. As shown in the original article, plant population at equilib-

rium is determined by the herbivore characteristics:

V0¼ (dh/w), where dh is the intrinsic mortality rate of herbivores and w isthe herbivore consumption rate. The model further assumes that plant

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402 Nicolas Loeuille et al.

defences sv decrease this consumption rate w, while herbivore trait shincreases the investment of the herbivore in consumption, at the expense

of mortality costs. Incorporating such traits, one gets:

V 0¼ dh shð Þw sh,svð Þ :

Because defences decrease the consumption rate w, in terms of pheno-

type variation, selecting phenotypes sv that produce higher total yield V0

means selecting higher values of sv. We point out that this result is true

for many other models describing plant–herbivore interaction, provided that

they use prey-dependent functional responses for plant consumption and

that herbivore mortality is density-independent.

APPENDIX C. EFFECTS OF ENRICHMENT ON THECONTROL OF BIOMASS WITHIN A

TRI-TROPHIC FOOD CHAIN WHEN THEHERBIVORE EVOLVES

We extend the model of Loeuille and Loreau (2004) by adding a

predator level. This makes the model much more relevant to some

agricultural questions, as biological control exerted by enemies of herbivores

is primordial for a sustainable agriculture. The ecological dynamics of the

system follow the set of ordinary differential equations:

dN

dt¼ I� lnN þ 1� vvð ÞdvV þ 1� vhð ÞdhHþ 1� vp

� �dpP�gNV

dV

dt¼V gN � dv�wHð Þ

dH

dt¼H wV � dh�aPð Þ

dP

dt¼P aH�dp� �

8>>>>>>>>>>>><>>>>>>>>>>>>:

ðC1Þwhere variables N, V, H and P correspond to the pool of limiting inorganic

nutrient, plant biomass, herbivore biomass and predator biomass, respec-

tively. All compartments are expressed in mass units of limiting nutrient.

I corresponds to the mass of nutrient input per unit of time in the system,

ln is the diffusion rate of the nutrient out of the system, vx is the proportion

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403Eco-Evolutionary Dynamics in Agroecosystems

of the nutrient of species x that is recycled within the system, dx is the turn-

over rate of the biomass of species x (which combines the basic death rate of

individuals, but also nutrient excreted by individuals, as well as dead tissues)

and g, w, a represent the individual consumption rates of plants, herbivores

and predators, respectively.

By setting the right side of Eq. (C1) to zero it is possible to determine the

coexistence equilibrium. This equilibrium reads:

N 0¼ wH0þ dv

g

V 0¼ a Ig�dvlnð Þþdpdhg vp� vh� ��wdpln

gwdpvpþ gadvvv

H0¼ dp

a

P0¼ wV 0� dh

a

ðC2Þ

Here exponent 0 stands for ‘ecological equilibrium’. It is possible to show

that, as soon as nutrient input is sufficient to maintain the coexistence equi-

librium, that is to say, as soon as P0>0, this coexistence equilibrium is stable.

The critical value of nutrient input needed to maintain the full food chain is:

I >wdplnþ gdhdpvh

gaþ dv

ln

gþ dhvh

w

� �ðC3Þ

Once this value is reached, further enrichment affects the equilibrium

biomass of the food chain as:

@N 0

@I¼ 0

@V 0

@I¼ aadvvvþwdpvp

@H0

@I¼ 0

@P0

@I¼ wadvvvþwdpvp

ðC4Þ

Equation (C4) is simply obtained by differentiating Eq. (C2) with respect

to nutrient input rates.

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404 Nicolas Loeuille et al.

Equation (C4) indicates that nutrient enrichment benefits plant and

predator biomass, while insect pests (here herbivores) are controlled by their

enemies. Therefore, the ecological dynamics of the system produces a desir-

able outcome from an agricultural point of view, with the cultivated organ-

ism benefitting from the enrichment and enemies are kept in check through

biological control. Such variations of the nutrient stocks are completely con-

sistent with those uncovered in previous works on food chains (Loeuille and

Loreau, 2004; Oksanen et al., 1981).

Now, we consider that herbivores evolve. We assume a trait x that

corresponds to the investment of herbivores in defences against their

predators. We voluntarily keep this trait as general as possible. It may

simply be some toxicity or morphologies that would help the herbivore

to escape predation, but also behaviours that would allow herbivores to

escape their predators (vigilance, use of alternative habitats, etc.). The

only assumption we make is that evolution towards higher values of

this trait decreases the growth or reproduction of the herbivore. In mathe-

matical terms, we assume that trait x can be any real value, and it affects

simultaneously w and a such that both are now decreasing functions of

the trait. That is:

x2Rw0 xð Þ< 0

a0 xð Þ< 0

ðC5Þ

Again, to maintain the desired level of generality, we do not explicit

functions a and w so that the following results do not depend on a priori fixedtrade-off shapes. We study the evolution of x using adaptive dynamics (see

Appendix A for the details of this technique). Individual fitness readily

emerge from population dynamics (Eq. C1):

W xm,xð Þ¼ 1

Hm

dHm

dt

����Hm!0

¼ w xmð ÞV 0 xð Þ� dh�a xmð ÞP0 xð Þ ðC6Þ

Evolutionary singularities are therefore determined by:

@W xm,xð Þ@xm

����xm!x

¼ 0, w0 x�ð ÞV � �a0 x�ð ÞP� ¼ 0 ðC7Þ

In Eq. (C7), x� corresponds to the value of the phenotype at the evolu-

tionary singularity and V� and P� correspond to the plant and predator bio-

masses at this singular trait. Computing the exact position of x� requires

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405Eco-Evolutionary Dynamics in Agroecosystems

defining more precisely the trade-off, that is to say, to define precisely the

functions a(x) and w(x). This is unnecessary for our argument, so we will

just assume that at least one value of x exists that satisfies Eq. (C7). Would

it not be the case, evolutionary dynamics would produce ever increasing or

ever decreasing values of x. Eventually, such runaway dynamics would

induce the extinction of the predator (evolutionary murder) or of the her-

bivore (evolutionary suicide).

Assuming that x� exists, it is now necessary to check whether evolution

will lead to an end at this point or not. The strategy cannot be invaded

provided:

@2W xm,xð Þ@xm2

����xm!x!x�

< 0, w00 x�ð ÞV � �a00 x�ð ÞP� < 0 ðC8Þ

The convergence condition can be determined from the following

condition:

@2W xm,xð Þ@xm2

����xm!x!x�

þ@2W xm,xð Þ@xm@x

����xm!x!x�

< 0

, w00 x�ð ÞV � �a00 x�ð ÞP� þa0 x�ð Þ P�@V@x

��x�

V � �@P

@x

����x�

!< 0 ðC9Þ

If the evolutionary singularity satisfies Eqs. (C8) and (C9) simulta-

neously, then the evolutionary singularity is called a CSS. Selection then

drives x towards the singular strategy x� and eventually stops at this point.

Note that conditions (C8) and (C9) are not necessarily fulfilled. The

trade-off shape and density-dependent effects of the trait x play a critical role

in that respect. Then, it is possible that evolution runs away from the strat-

egy. Alternatively, it is possible that evolution leads to the singularity, but

that it is invasible, so that a diversification occurs through branching pro-

cesses. These possibilities are interesting, but we can only understand

how enrichment affects the evolutionary equilibrium when this equilibrium

exists and is reached, so that we focus only on CSS cases. More complicated

dynamics are beyond the scope of the present analysis.

So we assume that conditions (C8) and (C9) are met. Phenotype x then

eventually stabilizes at x�, defined by Eq. (C7), and we disturb the system by

increasing nutrient inputs. Again, differentiating equilibrium condi-

tions (C2) together with the evolutionary equilibrium condition (C7),

one gets the following variations:

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406 Nicolas Loeuille et al.

@N�

@I¼�a0H�dh

gaV �@x�

@I

@V �

@I¼ aV � w00V � �a00P�ð Þ

a0dh

@x�

@I

@H�

@I¼�a0H�

a@x�

@I

@P�

@I¼ wa@V �

@I

@x�

@I¼� a2a0gdhV �

a02dh2dp lnþ gvhV �ð Þ�a2gV � w00V � �a00P�ð Þ wdpvpþadvvv� �

ðC10Þ

Recalling from condition (C5) that a0<0, and from condition (C8) that

w00V��a00P�<0, and recalling that other parameters and variables of the

model are all positive, it is easy to see from the first four lines of Eq. (C10) that

N�,V�,H� andP� all vary in the samedirection asx�whennutrient input rate Iis increased. Again from inequalities (C5) and (C8), it is easy to see that the

right-hand side of the last line of Eq. (C10) is positive, so that x� increaseswhen I increases. It follows from all these considerations that when nutrient

enrichment takes place, all biomasses increase, regardless of the trophic level,

and that the herbivore trait of investment in defence increases too. Therefore,

while the herbivore pest is controlled by its predator from an ecological

dynamics point of view (see Eq.C4), it is no longer controlledwhenwe allow

evolutionof defences for this herbivore. Its populationdoes increase through-

out enrichment (Eq. C10), due to this evolutionary effect. The combination

of nutrient enrichment and of the evolutionary process decreases the biolog-

ical control exerted by herbivores.Next to the qualitative analysis given here,

an example of such an effect is displayed on Fig. 6.3 in Box 6.3. To realize this

figure, we had to fix a trade-off shape. We assumed that:

a xð Þ¼ a0e�kax

w xð Þ¼ w0e�kwx

ðC11Þ

Trade-off Eqs. (C11) have the advantage that, by varying ka and kw one

can get a linear, convex or concave relationship between a and w, so that thecomplete variety of evolutionary outcomes are possible, including CSSs.

Parameter values we use for the figure are the following: ln¼0.06,

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407Eco-Evolutionary Dynamics in Agroecosystems

vv¼0.13, vp¼0.3, vh¼0.3, dv¼0.95, dp¼0.75, dh¼0.4, g¼2.7, w0¼1.4,

a0¼4.5, kw¼0.12, ka¼0.13.

APPENDIX D. ECO-EVOLUTIONARY DYNAMICS INA PLANT–HERBIVORE–PREDATOR

CONFRONTED TO INSECTICIDES

Using ordinary differential equations we study the influence of insec-

ticides on a three-species community as the following:

1

V

dV

dt¼ r� aV �wH D1ð Þ

1

H

dH

dt¼ cwV �aP�dh� lj D2ð Þ

1

P

dP

dt¼ gaH�dp� lh D3ð Þ

8>>>>>>>><>>>>>>>>:

The model describes a community composed of one plant V, one her-

bivoreH, and one predator P. The plant dynamics (D3) is constrained by the

intrinsic growth rate (r) and intra-specific competition a. Hence, in the

absence of the herbivore, the plant has a logistic growth. We used Holling

type I functional responses to model the consumption interactions, w being

the per capita attack rate of herbivores, c the energy conversion efficiency of

the herbivore, a the per capita attack rate of predators, and g the conversion

efficiency of this predator. Herbivore and predator dynamics also decrease

through interaction-independent death rates (respectively, dh and dp) and

through an additional mortality term, due to the action of insecticides.

The parameter l represents the intensity of insecticide application and

j and h correspond to the sensitivities of herbivores and predators, respec-

tively. The two populations are therefore confronted to the same perturba-

tion, to which they respond differently because of their specific sensitivities.

By setting the time derivatives in Eqs. (D1), (D2) and (D3) to zero, the

equilibrium points of the three-species model can be fully determined. The

plant–herbivore–predator model has four equilibrium points. Existence

condition requires that species densities at equilibrium are positive. We also

assess the stability of these points, with the Routh–Hurwitz criterion.

The trivial equilibrium S1¼ (V0¼0; H0¼0; P0¼0) always exists, but it

is unstable when r>0, a situation that is most often true given that the plant

is cultivated.

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408 Nicolas Loeuille et al.

The plant equilibrium point S2¼ (V0¼ r/a; H0¼0; P0¼0) always exists

and it is stable when the perturbation by insecticides is strong enough to pre-

vent the herbivore to invade the system. This requires:

a ljþ dhð Þ> wcr ðD4ÞThe third equilibrium point allows the coexistence of plant and herbi-

vores. At this equilibrium S3, plants and herbivores have the following den-

sity equations:

V 0¼ ljþdh

wcðD5Þ

H0¼�a ljþdhð Þþwcrw2c

ðD6Þ

Out of Eq. (D5) one clearly sees that the plant biomass is always positive.

Therefore, existence conditions are based on the positivity of herbivore

biomass. This requires that:

a ljþ dhð Þ< wcr ðD7ÞThe Routh–Hurwitz criterion shows that this plant–herbivore equilib-

rium point is stable whenmortality incurred by predator population through

pesticides is high:

aga dhþ ljð Þ> w �w dpþhl� �þagr

� �c ðD8Þ

Finally, it is possible that all three species coexist at one equilibrium,

where density equations are

V 0¼�w dpþhl� �þagr

agaðD9Þ

H0¼ dpþhl

agðD10Þ

P0¼ w �w dpþhl� �þagr

� �c�aga dhþ ljð Þ

a2gaðD11Þ

All the three species have positive densities when predator biomass (D11)

is positive. This is ensured by the following condition:

aga dhþ ljð Þ< w �w dpþhl� �þagr

� �c ðD12Þ

The Routh–Hurwitz criterion for a system of three ordinary differential

equations shows that the system is stable when condition (D12) is satisfied.

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409Eco-Evolutionary Dynamics in Agroecosystems

We now assume to study how these ecological equilibriums change

when more insecticides are used. By computing the partial derivatives of

equilibrium equations with respect to l, we can solve analytically how the

densities of the different species will vary with an increase in perturbation

intensity l.

When the community does not contain the predator, the partial deriv-

atives of Eqs. (D5) and (D6) with respect to l are

@V 0

@l¼ j

wcðD13Þ

@H0

@l¼� aj

w2cðD14Þ

Thus, when the intensity of perturbation increases, plant biomass

increases while herbivores are negatively affected.

We similarly study changes in the full coexistence equilibrium (Eqs. D9–

D11):

@V 0

@l¼� wh

agaðD15Þ

@H0

@l¼ h

agðD16Þ

@P0

@l¼�w2hcþagaj

a2gaðD17Þ

At the coexistence equilibrium, the plants and predator biomasses

decrease with an increase in insecticide use, while herbivores increase.

Comparisons of Eqs. (D13 and D14) and of Eqs. (D15–D17) highlight

the fact that the efficiency of pesticide use strongly varies with the number of

trophic levels. When the system is composed of two levels, then using pes-

ticide yields the intended result, namely the decrease of the pest population

and the increase in yield. When the system is composed of three trophic

levels and when the top trophic level is sensitive to insecticide, using them

erodes the biological control of pests and has adverse effects from an agricul-

tural point of view.

Now that we have obtained the effects of insecticide disturbances on the

ecological equilibrium and highlighted that qualitative variations are depen-

dent on the number of trophic levels, we study how evolution of agricultural

pests (here, herbivores) change such qualitative outcomes.

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410 Nicolas Loeuille et al.

As pointed out in the main text, the emergence of resistance in pests is

often associated with fitness costs (Bourguet et al. 2004; Carriere et al., 1994;

Gassmann et al., 2009). Particularly, studies have shown that resistance most

often decrease the allocation in growth or reproduction (Bourguet et al.

2004; Carriere et al., 1994). To consider this trade-off, we add in our

plant–herbivore–predator community a sensitivity trait for the herbivore

that affects both its mortality due to insecticides and its reproduction rate.

The model becomes

1

V

dV

dt¼ r� aV �wH D18ð Þ

1

H

dH

dt¼ c sð ÞwV �aP� dh� l sð Þj D19ð Þ

1

P

dP

dt¼ gaH� dp� lh D20ð Þ

8>>>>>>>><>>>>>>>>:

The two functions j(s) and c(s) are positive, increasing function of s. The

lower the value of s, the more resistant the pest is, but the higher the repro-

duction cost. We choose exponential functions for c and j, because they sat-

isfy the different assumptions and also for mathematical convenience:

c sð Þ¼ c0 exp vsð Þ ðD21Þj sð Þ¼ j0 exp zsð Þ ðD22Þ

The ratio v/z controls the shape of the trade-off. By varying their relative

values, we can have a large panel of trade-off shapes, from convexity to con-

cavity. When the trait of interest and the associated trade-off are set, we use

adaptive dynamics to determine the resulting eco-evolutionary dynamics

(see Appendix A for an introduction of this technique).

We first study the casewhere only plants andherbivores coexist.The fitness

of a rare mutant s0 in a resident population of trait s can then be determined:

W1 s0, sð Þ¼ 1

dHm

dHm

dt

����Hm!0

¼ c s0ð Þ wV 0 sð Þ� dh� lj s0ð Þ ðD23Þ

withV0¼ (lj(s)þm)/ac(s), the density of the plant at the equilibrium fixed by

the resident trait.

The associated fitness gradient that determines the direction of change of

the trait is

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411Eco-Evolutionary Dynamics in Agroecosystems

@W1 s0, sð Þ@s0

����s0!s

¼ vdhþ lj sð Þ v�zð Þ ðD24Þ

From Eq. (D24), two qualitative possibilities emerge, depending on the

trade-off shapes. If v�z the fitness gradient is always positive. The sensitivity

of herbivore s always increases during the evolution. When v<z, the fitness

gradient canbepositive or negative dependingon the resident trait value. Sin-

gular strategies may then be obtained, at which the gradient of fitness is null.

Such a singular strategy is given by the following equation:

s�1 ¼1

zlog

vdh

j0l z� vð Þ� �

ðD25Þ

As inAppendicesA andC,we study the invasibility and convergence char-

acteristics of this singular strategy. Recalling that we have v<z, we obtain:

@2W1 s0, sð Þ@s02

����s0!s!s�

1

¼ dhvz v�zð Þz� v

< 0 ðD26Þ

Because this second derivative is negative, the singular strategy cannot be

invaded by surrounding mutants.

It can be easily shown that:

@2W1 s0, sð Þ@s0@s

����s0!s!s�

1

¼ 0 ðD27Þ

The sum of Eqs. (D26) and (D27) is always negative, which ensures that

the singular strategy is convergent, that is, evolution drives the system

towards this point. The evolutionary singularity is therefore a CSS.

We now characterize the evolutionary dynamics when all three species

coexist. The invasion fitness of a mutant herbivore is:

W2 s0, sð Þ¼ wV 0c s0ð Þ�aP0 sð Þ� dh� lj s0ð Þ ðD28ÞThe fitness gradient becomes

@W2 s0, sð Þ@s0

����s0!s

¼ wV 0c0 sð Þ� lj0 sð Þ ðD29Þ

The singular strategy, at which the gradient vanishes is

s�2 ¼1

v�zlog

j0lz

c0wvV 0

� �ðD30Þ

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412 Nicolas Loeuille et al.

Note that, in Eq. (D28), plant biomass is not dependent on the resident

trait. Therefore, it is easy to see that @2W(s0, s)/@s0 @s is equal to zero. Then,

as in the plant–herbivore case, characteristics of the evolutionary singularity

only depend on the second derivative with respect to s0:

@2W2 s0, sð Þ@s02

����s0!s!s�

2

¼ v2wV 0c s�2� ��z2lj s�2

� � ðD31Þ

Inserting Eq. (D30) in Eq. (D31) shows that it is a decreasing function of

z and that it vanishes for z¼v. Because at the evolutionary

singularity, the system is also at equilibrium from an ecological point of

view, wV 0c(s�) ¼aP0(s�)þdhþ lj(s�). Equation (D31) becomes

@2W2 s0, sð Þ@s02

����s0!s!s�

2

¼ v2 aP0 s�2� �þ dh

� �þ v2�z2� �

lj s�2� � ðD32Þ

Equation (D32) is positive for v>z.

Therefore, the second partial derivative of invasive fitness with respect to

s0 is positive for v>z, vanishes for v¼z and is negative for v<z. We con-

clude from this that the singular strategy in this model is a repellor when the

trade-off is convex (v>z) and a CSS when the trade-off is concave (v<z).

The repellor situation leads to a runaway dynamics on which we cannot eas-

ily study the influence of increased pesticides. For the CSS cases, on the

other hand, evolution eventually settles the trait at s2� and we can study

the effects of increased pesticide use by linearizing around this equilibrium

situation.

To study the influence of increased pesticide use, we differentiate

equations of the singular strategies s1� and s2� (Eqs. D25 and D30) regarding

pesticide use parameter l.

When the system only contains plants and herbivores, effects of pesticide

use on the position of the singular strategy are determined by the following

equation:

@s1�

@l¼� 1

lzðD33Þ

In such instances, the singular strategy always decreases when pesticide

use increases. From an agricultural point of view, it means that we expect

evolution of further resistance in all instances when more insecticide is

sprayed.

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413Eco-Evolutionary Dynamics in Agroecosystems

When the three species coexist, variations in the singular strategy with

pesticide use are determined from the following equations:

@s�2@l

¼ wdp�agrl w dpþhl� ��agr

� �v�zð Þ ðD34Þ

@s2�

@h¼ wl

�w dpþhl� �þagr

� �v�zð Þ ðD35Þ

Note that again, given condition of coexistence (D12), when the singular

strategy is a CSS, it decreases with l, meaning that increased pesticide use

allows for evolution of higher levels of resistance. The sensitivity of the pred-

ator also affects on the singular strategy. The singular strategy decreases with

predator sensitivity h. In agricultural terms, it clearly indicates that sensitive

predators allows for higher levels of resistance in the pest crop.

APPENDIX E. EVOLUTION OF SPECIALIZATION RATEOF THE PEST AND ITS ECOLOGICAL

CONSEQUENCES

We tailor the model of Leibold (1996) to fit the agricultural scenario of

an insect-pest feeding on two plant types, one weed and one crop, which in

their turn compete over a common soil resource. The ecological dynamics

of this system follow the set of ordinary differential equations:

dH

dt¼H hc xcþxοc

� �Vcþhw xwþxοw

� �Vw�dh

dVc

dt¼Vc vcgcR� wcþwoc

� �H�dc

dVw

dt¼Vw vwgwR� wwþwow

� �H� dw

dN

dt¼ I � lN � gcNVc� gwNVw

ðE1Þ

Where wio and wi (i¼c, w) denote the plant and herbivore-dependent parts of

the plant–herbivore interaction; when the plant is maximally defended,

wio¼0. Note that we suppose these density-dependent effects are additive.

The herbivore-dependent part wi can also be interpreted as degree of special-ization on the target plant i. Other variables and parameters have been

defined in Appendix C. The only additional parameters appearing in this

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414 Nicolas Loeuille et al.

model are the conversion efficiencies in the functional responses of the pest

and the two plant species (hc, hw, vc, vw).

Here, wemake some fairly realistic and general assumptions about the crop

and the weed species. First, crop most often requires fertilization to grow and

survive, so that we assume the crop is poor competitor for the resource com-

pared to the weed (i.e.,dw

vwgw<

dc

vcgc). We also assume that the crop species is

relatively less vulnerable to herbivores owing to its enhanced herbivore resis-

tance favoured during breeding efforts (wco<<ww

o so that wwþwwo >wcþwc

o).

In other words, while the crop is mostly limited by the resource, the weed is

mostly limited by the pest. Theoretically, the stable coexistence of the two

competing plant species is assured when the following two conditions are

met: (i) the ratio of resource effects and herbivore effects on the per capita

growth of the crop species is higher than the same ratio of effects on the

per capita growth of the weed (i.e.,vcgcXc

>vwgwXw

, where Xi¼wiþwiο),

(ii) the ratio of weed impacts on herbivore and resource growth is higher than

the same ratio of crop impacts (i.e.,hwXwH

gwR>hcXcH

gcR) (Leibold, 1996).

The coexistence equilibrium of the set of Eqs. (E1) reads as follows:

N0¼dc

Xc

� dw

XwvcgcXc

� vwgwXw

0BB@

1CCAV 0

c ¼dh�hwXw

gw

I� lN0

N0

� �

gchcXc

gc�hwXw

gw

� � V 0w¼

dh�hcXc

gc

I� lN0

N0

� �

gwhwXw

gw�hcXc

gc

� �

H0¼ vcgcXc

vwgwXw

dc

vcgc� dw

vwgwvcgcXc

� vwgwXw

0BB@

1CCA ðE2Þ

Note that the equilibrium biomass of the crop (weed) increases

(decreases) linearly with the resource input rate I (i.e.,@V 0

w

@I< 0,

@V 0c

@I> 0)

whereas the equilibrium biomasses of the resource and the pest remain

unchanged (@H0

@I,@R0

@I¼ 0).

Now, we consider that the pest evolves. In specific, we assume that the

degree of specialization on the crop (wc) coevolves with the degree of spe-

cialization on the weed (ww) in a traded-off manner. This trade-off can be

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415Eco-Evolutionary Dynamics in Agroecosystems

explained in terms of energetic limitations: the energy spent in searching and

handling of one plant type comes to a cost for searching and handling of the

other one. Again we assume that a trait x affects both traits and ww such that

w0c xð Þ> 0, then w0w xð Þ< 0.

Next, we follow the methodological steps of adaptive dynamics, as intro-

duced in appendix A. First, we define the relative fitness of a rare variant xmin a population x as follows:

WH xm,xð Þ¼ hc wc xmð Þþwοc� �

V 0c þhw ww xmð Þþwοw

� �V 0w� dh ðE3Þ

The evolutionary singularity of the above fitness function is obtained by

setting the first derivative of the above fitness function to zero:

@Wh xm,xð Þ@xm

����xm¼x¼x�

¼ 0, hcw0c x�ð ÞV �c þhww0w x�ð ÞV �

w¼ 0 ðE4Þ

where x� denotes the singular strategy and Vc�,Vw

� correspond to the equi-librium biomass of the crop and the weed at the singularity, respectively.

Then, the invasibility of the singular strategy x� can be determined by the

following condition:

@2Wh xm,xð Þ@x2m

����xm¼x¼x�

< 0, hcw00c x�ð ÞV �c þhww00w x�ð ÞV �

w< 0 ðE5Þ

The convergence stability of the singular strategy can be determined

from the following condition:

@2Wh xm,xð Þ@x2m

����xm¼x¼x�

þ@2Wh xm,xð Þ@xm@x

����xm¼x¼x�

< 0

, hcw00c x�ð ÞV �c þhww00w x�ð ÞV �

wþhc@V �

c

@x

����x��V �

c

V �w

@V �w

@x

����x�

� �< 0

ðE6Þ

It follows that the singular strategy represents a CSS when both special-

ization rates saturate with increasing x (i.e., w00c xð Þ,w00w xð Þ are negative). We

focus on the CSS case and again look at the adaptive changes generated in the

equilibrium biomasses of the model compartments when we perturb the sys-

tem by changing input rate I (e.g., due to fertilization). Noting, as in the

other appendices, the ecological equilibrium by 0 and the evolutionary equi-

librium by �, the adaptive responses of species biomass to the above pertur-

bation were calculated to have the following signs (explicit formulas are not

provided because of their complexity):

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416 Nicolas Loeuille et al.

@H�

@I> 0,

@V 0c

@I>@V �

c

@I> 0,

@V 0w

@I<@V �

w

@I< 0,

@R�

@I> 0 ðE7Þ

and

@H�

@dw< 0,

@V 0c

@dw>@V �

c

@dw> 0,

@V 0w

@dw<@V �

w

@dw< 0,

@R�

@dw< 0

Several things are worth noticing in the above line of inequalities: first,

the perturbation parameter I generates a positive effect on trait x, which

translates to a positive effect on the trait wc and a negative one on the trait

ww. Second, the ecological effect of a change in I on plant biomasses is mod-

erated by the evolution of specialization (See Fig. 6.6). Therefore, consistent

with results detailed in other appendices, these last results suggest that using

purely ecological models may give over-optimistic results compared to the

complete eco-evolutionary feedbacks. Third, the evolution of specialization

enforces the top-down control of the herbivore, which results in a positive

response of the herbivore’s biomass to fertilization.

To realize Fig. 6.6, we had to fix a trade-off shape. We assumed that:

wc xð Þ¼ wmaxc x1=z

ww xð Þ¼ wmaxw 1�xð Þ1=z, wherex2 0,1½ � ðE8Þ

The parameter z measures the strength of the trade-off; when z>1 the

trade-off is weak (w00c xð Þ,w00w xð Þ are negative) and when z<1 the trade-off is

strong (w00c xð Þ,w00w xð Þ are positive).Parameter values we use for the Fig. 6.6 are the following: hc¼0.5,

hw¼0.7, wcmax¼1, wc

0¼0.01, wwmax¼1, ww

0¼0.1, dh¼0.1, dc¼0.4,

dw¼0.2, vc¼0.5, vw¼1, gc¼0.6, gw¼0.3, l¼0.1, z¼1.5.

GLOSSARYEcological network a complex set of species linked by interactions of different types:

direct/indirect, trophic/non-trophic, mutualistic/antagonistic. The complexity charac-

terizing an ecological network constrains species ecological dynamics and adaptive

responses to disturbances.

Fitness (individual) number of reproducing offspring left by an individual during its

lifetime.

Adaptability ability of a population to cope with environmental variations through genetic

change (natural selection, gene flows), phenotypic plasticity or behavioural change.

Natural selection a gradual, non-random process by which phenotypic/genotypic traits

change in frequency as a function of fitness differences among their bearers. Important

features of selection include:

Dimensionality number of selective pressures operating simultaneously,

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417Eco-Evolutionary Dynamics in Agroecosystems

Intensity difference between maximum and minimum fitness within the population. It

depends on environmental changes/fluctuation, changes in the ecological context, gene

flow, etc.

Group selection (in the case of agriculture, Artificial Group Selection) selection

based on the fitness of the group rather than individual fitness; this may result in fixation

of traits disadvantageous to the individual itself, provided there is some heritability of the

group property under selection.

Artificial selection human-driven selection of organisms favouring those with desirable

characteristics, for example, for cultivation and use. These characteristics may be nega-

tively linked to organism fitness.

Adaptive change/response a phenotypic change (with a genetic or non-genetic basis) that

improves the fitness of individuals relative to average fitness within the population; it can

be limited by allocation or ecological trade-offs.

Allocation trade-off beneficial changes in one trait cause detrimental changes in another

due to energetic or time constraints within an individual.

Ecological trade-off beneficial changes in a species response to one interaction incur costs

from another interaction (e.g., plant defences against herbivores may have detrimental

effects on pollination); the higher the complexity of an ecological network the more

important these trade-offs are.

Domestication outcome of the artificial selection process linked with cultivation by

humans of plants and animals.

Diffuse co-evolution allelic/phenotypic changes occurring within the frame of an ecolog-

ical network, that is, species evolve in response to a number of other species of the net-

work, each of which is also evolving in response to another set of species (Janzen, 1980).

Contemporary (or rapid) evolution allelic/phenotypic change occurring in natural eco-

systems on a short time frame

Eco-evolutionary feedback situation in which the ecological context constrains natural

selection, while resulting evolution affects the ecological context through changes in spe-

cies density, species interactions or the environment.

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