Responses to landscape changes

ORIGIN AND CAUSES OF LANDSCAPE CHANGES

Despite our tendency to consider Earth as static, it is actually a dynamic and ever-changing planet. Wind, water, and ice erode and shape the land. Natural fires, storms, flooding, dryness also affect habitats. Volcanic activity and earthquakes alter the landscape in a dramatic and often violent manner. And on a much longer timescale, the movement of earth’s plates slowly reconfigures oceans and continents. 

Humans have also long modified landscapes through controlled burns. During the Paleolithic and Mesolithic ages, fire was used extensively for what has been termed “fire-stick farming” (Bird et al. 2008). This term implies using fire for a variety of reasons: clearing ground for human habitats, facilitating travel, killing vermin, hunting, regenerating plant food sources for both humans and livestock, and even warfare among tribes. These land-management practices had profound impacts not only on fire regimes but also on the landscape vegetation pattern and biodiversity. Later, about 12'000 years ago, our hunter-gatherer ancestors began trying their hand at farming. This marks the start of the Neolithic revolution, which allowed agriculture to spread, with as corollaries deforestation. According to Naveh and Dan (1973), human impact has exerted sustained, direct effects on Mediterranean ecosystems since at least 50,000 years ago, but a true ‘revolution’ began about 10,000 years ago, when hunters in the Near and Middle East began to produce their own food supply, thus laying the foundations for the domestication of plants and animals (Harris, 1998; Meadow, 1998; Miller, 1992; Vavilov, 1935; Zeder and Hesse, 2000). 

But the growth of technology for the last centuries has increased our ability to change a natural landscape. Many human activities increase the rate at which natural processes shape the landscape. Among anthropogenic causes of landscape changes and habitat degradation, we can cite: commercial logging, mining, clearance for agriculture and livestock farming, artificial fires, urbanization, transport infrastructures, dams, pipelines, climate changes, pollution... The effects of global change due to the rapid intensification of anthropogenic activities on natural habitats are thus became more and more evident (Parmesan, 2006; Travis, 2003). 

Forest conversion is the largest cause of global deforestation today. It is defined as the clearing of natural forests (deforestation) to use the land for another purpose, often agricultural (growing crops like palm oil or creating pasture for cattle), but also for mines, infrastructure or urbanisation. By the beginning of the 21st century, 75% of the Earth ice-free land had been modified by humans in some degree (Watson et al., 2016). Approximately two-third of these human-altered areas have been subject to the complete conversion of previously existing native vegetation. Between 2000 and 2015, global rates of deforestation of native forests ranged between 5,770 and 10,483 million ha/year, leading to a global loss of 124.8 million ha of  native forests during this period (Keenan et al., 2015), an area roughly twice the size of France. Such vast conversion has led to a substantial reduction in the areas considered to remain wild, i.e., with negligible human impacts. In 2016, only 23.2% (3.01 billion ha) of the world’s terrestrial area meets this definition (Watson et al., 2016), and these wilderness areas are scattered around the globe. While some remain considerably large, such as the Amazon forest or the West Siberian Taiga (Watson et al., 2016), land conversion has led to a vast swath of fragmented native vegetation across most biogeographic realms and phytogeographical domains. 

Between 1957 and 2012 in Europe, the road network increased from 146.7 to 421.8 million km (Stelder 2014), a 2.9-fold increase. Although 80% of the Earth's surface is currently without roads, the fragmentation of the environment is marked with the cutting of this surface into approximately 600,000 patches, more than half of which are less than 1 km² and only 7% are larger than 100 km² (Ibisch et al., 2016).

LANDSCAPE CHANGES AS A KEY DRIVER OF BIODIVERSITY LOSS

Landscape and habitat changes have been recognized throughout the world as a key issue affecting biodiversity (Myers et al. 2000; Ganzhorn et al. 2003). It is often considered as the most pervasive threat on biodiversity, impacting 86% of threatened mammals, 86% of threatened birds, 88% of threatened amphibians (IUCN).

Some regions of our planet are particularly impacted by habitat loss and fragmentation. In particular, biodiversity hotspots are biogeographic regions where exceptional concentrations of endemic species are undergoing exceptional loss of habitat (Myers 2000). To qualify as a biodiversity hotspot, a region must contain at least 0.5% or 1,500 species of vascular plants as endemics, and it has to have lost at least 75% of its primary vegetation. Around the world, 36 areas qualify under this definition. These sites support nearly 60% of the world's plant, bird, mammal, reptile, and amphibian species, with a very high share of those species as endemics. Some of these hotspots support up to 15,000 endemic plant species and some have lost up to 95% of their natural habitat. Biodiversity hotspots host their diverse ecosystems on just 2.4% of the planet's surface, however, the area defined as hotspots covers a much larger proportion of the land. Overall, the current hotspots cover more than 15.7% of the land surface area, but have lost around 85% of their habitat. This loss of habitat explains why approximately 60% of the world's terrestrial life lives on only 2.4% of the land surface area.

LANDSCAPE STRUCTURE AND CONNECTIVITY

In landscape ecology, a practical model to conceptualize and represent the landscape features is the 'corridor-patch-matrix' landscape model (Forman 1995). According to this model, the landscape consists of habitat patches, corridors connecting habitat patches and a surrounding matrix, which is unsuitable for the species in the habitats (Levin 2006). 

• The landscape is composed of a mosaic of patches (Urban et al., 1987). 

As explained here (http://www.umass.edu/landeco/research/fragstats/documents/Conceptual%20Background/Patch-Corridor-Matrix%20Model/Patch-Corridor-Matrix%20Model.htm), "From an ecological perspective, patches represent relatively discrete areas (spatial domain) or periods (temporal domain) of relatively homogeneous environmental conditions where the patch boundaries are distinguished by discontinuities in environmental character states from their surroundings of magnitudes that are perceived by or relevant to the organism or ecological phenomenon under consideration (Wiens 1976). From a strictly organism-centered view, patches may be defined as environmental units between which fitness prospects, or "quality", differ; although, in practice, patches may be more appropriately defined by nonrandom distribution of activity or resource utilization among environmental units, as recognized in the concept of "Grain Response". The patches are dynamic, i.e. changing in time and space, and are found at various spatial and temporal scales that vary according to the perceptions of each animal (Wiens, 1976, 1989; Wiens and Milne, 1989). Regardless of the basis for defining patches, a landscape does not contain a single patch mosaic, but contains a hierarchy of patch mosaics across a range of scales. From an organism-centered perspective, patches can be defined hierarchically in scales ranging between the grain and extent for the individual, deme, population, or range of each species.FPatch boundaries are artificially imposed and are in fact meaningful only when referenced to a particular scale (i.e., grain size and extent)."

• The matrix that separates patches is defined as the most common or connected landscape element type that generally plays the dominant role in landscape function (Forman & Gordon 1986).

As explained here (http://www.umass.edu/landeco/research/fragstats/documents/Conceptual%20Background/Patch-Corridor-Matrix%20Model/Patch-Corridor-Matrix%20Model.htm), "The designation of a matrix element is largely dependent upon the phenomenon under consideration. For example, in the study of geomorphological processes, the geological substrate may serve to define the matrix and patches; whereas, in the study of vertebrate populations, vegetation structure may serve to define the matrix and patches. In addition, what constitutes the matrix is dependent on the scale of investigation or management. For example, at a particular scale, mature forest may be the matrix with disturbance patches embedded within; whereas, at a coarser scale, agricultural land may be the matrix with mature forest patches embedded within."

• Corridors are linear landscape elements that can be defined on the basis of their structure or function. 

They can have either a anthropogenic origin (such as roads and tracks) or a natural origin (such as ridges, valley bottoms, rivers and forest edges).

Forman & Godron (1986) define corridors as “narrow strips of land which differ from the matrix on either side. Corridors may be isolated strips, but are usually attached to a patch of somewhat similar vegetation.” They focus on the structural aspects of the linear landscape element. As a consequence of their form and context, structural corridors may function as habitat, dispersal conduits, or barriers. 

As explained here (http://www.umass.edu/landeco/research/fragstats/documents/Conceptual%20Background/Patch-Corridor-Matrix%20Model/Patch-Corridor-Matrix%20Model.htm):

"Three different types of structural corridors exist: 

(1) line corridors, in which the width of the corridor is too narrow to allow for interior environmental conditions to develop; 

(2) strip corridors, in which the width of the corridor is wide enough to allow for interior conditions to develop; 

(3) stream corridors, which are a special category.

Corridors may also be defined on the basis of their function in the landscape. At least four major corridor functions have been recognized, as follows:

1.   Habitat Corridor.--Linear landscape element that ensure survival, natality, movement and gene flow between patches, and may provide either temporary or permanent habitat. Habitat corridors passively increase landscape connectivity for the focal organism(s).

2.   Facilitated Movement Corridor.–Linear landscape element that ensure survival, movement and gene flow, but not necessarily natality, between other habitat patches. Facilitated movement corridors actively increase landscape connectivity for the focal organism(s).

3.   Barrier or Filter Corridor.–Linear landscape element that prohibits (i.e., barrier) or differentially impedes (i.e., filter) the flow of energy, mineral nutrients, and/or species across (i.e., flows perpendicular to the length of the corridor). Barrier or filter corridors actively decrease matrix connectivity for the focal process.

4.   Source of Abiotic and Biotic Effects on the Surrounding Matrix.–Linear landscape element that modifies the inputs of energy, mineral nutrients, and/or species to the surrounding matrix and thereby effects the functioning of the surrounding matrix.

The function of the corridor will vary among organisms due to the differences in how organisms perceive and scale the environment."

Linear landscape features can also be a source of potential disturbance through associated human activity (e.g. Carricondo-Sanchez et al. 2020; Cole et al. 2015). Even when these structures can be crossed, they may be perceived by animals as a barrier and constitute as such behavioural barriers (Beyer et al, 2016). 

The possibility of movement between patches of the landscape mosaic is a function of landscape composition and configuration and the adaptation of organism behaviour to these two variables. This is what defines landscape connectivity (Taylor et al., 1993).

Connectivity assesses the extent to which a landscape facilitates or impedes ecological flows or functionality. The flows include interfragment movements of animals. 

We usually distinguish 2 types of connectivity (Moilanen and Nieminen 2002): 

• structural connectivity is defined by, e.g., the spatial pattern of the remaining habitat, the interfragment distance, and the presence of corridors, but does not incorporate the behavioral response of the individuals studied,

• functional connectivity is an estimate of the relationship between the landscape spatial pattern and the capacity of the species of interest to move through the landscape. Functional connectivity is species-specific.

HABITAT LOSS AND FRAGMENTATION

One of the major ways that landscape changes affects biodiversity is through habitat loss and fragmentation.

Habitat loss generally refers to the decrease in the spatial extent of natural habitat, whereas habitat fragmentation is often defined as a process during which a large expanse of habitat is transformed into a number of smaller patches of smaller total area, isolated. Habitat loss and fragmentation usually occur concurrently and are interrelated, both influencing biodiversity and ecological processes. 

The definition of habitat fragmentation implies four effects of the process of fragmentation on habitat pattern (Fahrig 2003): 

• reduction in habitat amount, 

• increase in number of habitat patches, 

• decrease in sizes of habitat patches, 

• increase in isolation of patches.

Habitat loss has usually a much stronger negative effect on biodiversity measures than habitat fragmentation per se (the “breaking apart” of habitat on biodiversity, that are independent of or in addition to the effects of habitat loss), which generally has weak effects that can be both positive and negative for the biodiversity (Fahrig, 2003). Some theoretical studies suggest that the effects of fragmentation per se should become apparent only at low levels of habitat amount, below approximately 20–30% habitat on the landscape (Fahrig 1998, Flather & Bevers 2002). 

Negative effects of fragmentation are likely due to two main causes (Fahrig 2003): 

• "First, fragmentation per se implies a larger number of smaller patches. At some point, each patch of habitat will be too small to sustain a local population or perhaps even an individual territory. Species that are unable to cross the non habitat portion of the landscape (the “matrix”) will be confined to a large number of too-small patches, ultimately reducing the overall population size and probability of persistence. 

• The second main cause of negative effects of fragmentation per se is negative edge effects; more fragmented landscapes contain more edge for a given amount of habitat. This can increase the probability of individuals leaving the habitat and entering the matrix. Overall the amount of time spent in the matrix will be larger in a more fragmented landscape, which may increase overall mortality rate and reduce overall reproductive rate of the population (Fahrig 2002). In addition, there are negative edge effects due to species interactions (e.g., increased predation on forest birds at forest edges; Chalfoun et al. 2002)." 

But positive effects of habitat fragmentation are also common (Fahrig 2003): 

• "One reason is that immigration rate is a function of the linear dimension of a habitat patch rather than the area of the patch. 

• Another one is that if habitat amount is held constant, increasing fragmentation per se actually implies smaller distances between patches."

IMPACT OF LANDSCAPE CHANGES ON BIODIVERSITY

Landscape change effects on biodiversity cover a range of response variables (Fahrig 2003), including direct measures of biodiversity, such as species richness (Findlay & Houlahan 1997, Gurd et al. 2001, Schmiegelow & Mönkkönen 2002, Steffan-Dewenter et al. 2002, Wettstein & Schmid 1999), population abundance and distribution (Best et al. 2001, Gibbs 1998, Guthery et al. 2001, Hanski et al. 1996, Hargis et al. 1999, Hinsley et al. 1995, Lande 1987, Sanchez-Zapata & Calvo 1999, Venier & Fahrig 1996) and genetic diversity  (Gibbs 2001); but also indirect measures of biodiversity and factors affecting biodiversity.

Landscape barriers and/or habitat fragmentation have marked effects on large-scale movements such as migration or dispersal (Schtickzelle and Baguette, 2003; Tucker et al., 2018; Turner, 1996), potentially limiting gene flow (Frankham et al.., 2002; Martínez et al., 2002) and impacting the survival of small populations (Turner, 1996; Wilcox and Murphy, 1985). In particular, road construction, agricultural development and dam construction create physical barriers for long-distance migratory species (Hardesty-Moore et al., 2018). For example, Harris et al (2009) reported that of the 24 large migratory ungulates they examined, six migrations were lost due to human infrastructure. However, landscape structures can also animal movements at smaller spatial scales (e.g. Leblond et al., 2010). An increasing number of daily movements are hindered or blocked by natural and/or anthropogenic landscape features (Dyer et al., 2002; Sánchez-Prieto et al., 2010). 

Changes in animal movements will affect in turn spatial ecology and in particular home range size, as well as habitat use and selection at different spatial and temporal scales (Johnson 1980):

• the first order selection, for which species distribution can be identified at the landscape scale,

• the second order selection, where the location of an individual's home range is known within the landscape, 

• the third order selection, which can determine selection within the home range based on fine-scale movement behavior,

• the fourth order selection, which involves investigations into how invasive species use or disturb habitats at the finest scale.

The impact of landscape structures and changes on individual movements results from the the interplay among four basic mechanistic components of organismal movement (Nathan et al. 2008): 

• the internal state (why move?), 

• motion (how to move?), 

• the navigation (when and where to move?) capacities of the individual, 

• the external factors affecting movement. 

Habitat loss and fragmentation can lead to a rapid decline in population abundance and inter-population gene flow, as well as an increase in genetic drift and inbreeding. These changes in microevolutionary forces can alter genetic diversity and structure, which in turn can affect the long-term evolutionary potential of populations and species and their risk of extinction (Keller & Waller 2002, Keyghobadi 2007). 

Landscape changes, by altering resource distribution (e.g., habitat quality, spatial configuration of patches), interspecific interactions (e.g., predator–prey and host–parasite dynamics, human disturbance), and sex (mate availability and inbreeding risk), are expected to affect various individual behaviours (including groupe size, diet, home range, social interactions, mating systems and tactics, movements, habitat use and selection, activity patterns) as well as physiological stress. 

These behavioural responses may impact individual fitness and population dynamics and genetics, and in turn, contribute to population viability and broad-scale patterns of distribution and abundance in fragmented landscapes (Banks et al. 2007). 

Habitat loss may negatively impact population growth rate (Bascompte et al. 2002), by affecting survival and breeding success (Kurki et al. 2000), as well as dispersal success (Bélisle et al. 2001, Pither & Taylor 1998, With & Crist 1995, With & King 1999). 

Landscape changes can affect trophic network, community structure and dynamics, and ecosystem functioning. 

In particular, habitat loss has been shown to affect community and ecosystem structure and dynamics, by reducing trophic chain length (Komonen et al. 2000), altering species interactions (Taylor & Merriam 1995 and in particular predation rate (Bergin et al. 2000, Hartley & Hunter 1998), and reducing the number of specialist, large-bodied species (Gibbs & Stanton 2001).

According to Haddad et al. (2015):

"Reduced fragment area and increased fragment isolation generally reduced abundance of birds, mammals, insects, and plants. This overall pattern emerged despite complex patterns of increases or declines in abundance of individual species with various proximate causes such as release from competition or predation, shifts in disturbance regimes, or alteration of abiotic factors. Reduced area, increased isolation, and increased proportion of edge habitat reduced seed predation and herbivory, whereas increased proportion of edge caused higher fledgling predation that had the effect of reducing bird fecundity (represented together as trophic dynamics). Perhaps because of reduced movement and abundance, the ability of species to persist was lower in smaller and more isolated fragments .

Fragmentation strongly reduced species richness of plants and animals (...), often changing the composition of entire communities. In tropical forests, reduced fragment size and increased proportion of edge habitat caused shifts in the physical environment that led to the loss of large and old trees in favor of pioneer trees, with subsequent impacts on the community composition of insects. In grasslands, fragment size also affected succession rate, such that increased light penetration and altered seed pools in smaller fragments impeded the rate of ecological succession relative to that of larger fragments.

Consistently, all aspects of fragmentation—reduced fragment area, increased isolation, and increased edge—had degrading effects on a disparate set of core ecosystem functions. Degraded functions included reduced carbon and nitrogen retention, productivity, and pollination.

The effects of fragmentation are mediated by variation in traits across species. More realistic predictions of community responses to fragmentation emerged after explicit consideration of species traits such as rarity and trophic level, dispersal mode, reproductive mode and life span, diet, and movement behavior."

Landscape changes can ultimately have long-term and progressive effects of fragmentation through: 

• extinction debt: temporal lags in species extinction in fragments. 

• immigration lag: small or isolated fragments are slower to accumulate species during community assembly.

• ecosystem function debt: manifested both as delayed changes in nutrient cycling and as changes to plant and consumer biomass.

As explained by Haddad et al. (2015): "Functional debts can result from biodiversity loss, as when loss of nutrients and reduction in decomposition are caused by simplification of food webs. Alternatively, the impact is exhibited through pathways whereby fragmentation changes biotic (for example, tree density in successional systems) or abiotic conditions (for example, light regimes or humidity) in ways that alter and potentially impair ecosystem function [for example, biomass collapse in fragments; altered nitrogen and carbon soil dynamics]."

Species specific response to landscape changes

The consequences for animals are manifold and complex (Fahrig, 2003) and can be positive or negative, depending on the behavioural response to the physical landscape structure (Dickson & Beier, 2002; Poessel et al., 2014), as well as species-specific habitat preferences and sensitivity to habitat heterogeneity. 

For example, for species that preferentially use ecotones, a given level of habitat fragmentation may be beneficial in terms of resource availability and diversity (Torres, Carvalho, Panzacchi, Linnell, & Fonseca, 2011). Some species may also benefit from habitat fragmentation through greater local heterogeneity in resource patches with complementary functions (e.g. food resources and shelter in agricultural landscapes for roe deer; Hewison et al., 2001) or due to a higher density of corridors used for movement (e.g. roads or tracks for wolves; James and Stuart-Smith, 2000). In contrast, the dense network of linear landscape features in highly fragmented habitats may be detrimental when these features act as physical and/or behavioural barriers (e.g. caribou, Leblond et al., 2013; migratory herbivores, Seidler et al., 2015). 

Linear landscape features are able to have very different roles on the movements of animals and on their spatial distribution, linked both to the environmental context (scale and level of heterogeneity of the landscape), the nature of the anthropogenic disturbance (potentially lethal or not, intensity, frequency, predictability), individual characteristics (e.g. personality), and finally, life history traits (e.g. predation avoidance behaviors, diet types and plasticity) and ecological traits of the species (e.g. level of gregariousness of species).

Hence, importantly, the definitions of fragment and connectivity should be species-specific, and a function of the species’ habitat requirements and dispersal capacity (Arroyo-Rodriguez & Mandujano 2009).

In the context of global changes stemming from an ongoing and ever growing human footprint on the planet, it is therefore crucial and urgent to better understand how animal populations respond to landscape changes and persist or not in human-modified environments. 

A major issue for wildlife conservation and management is to understand how these landscape effects, for example, habitat loss and fragmentation, will affect individual behaviours, species distribution, population dynamics and population genetics, and in turn, species interactions, community and ecosystem structure, dynamics and functioning.

Lemur project

Madagascar, known for its extreme biological diversity and high level of endemism, has lost 90% of its "original" forest cover and many forest habitats are now fragmented. Lemurs, endemic primates dependent on forest environments, are particularly vulnerable to these changes. 

My 2-year postdoctoral fellowship at IGC (Portugal) thus aimed, in the framework of a major project on population genetics and conservation of lemurs in Madagascar coordinated by Lounès Chikhi, to study the interspecific diversity of genetic responses to habitat loss and fragmentation of several lemur species in northern Madagascar. 

More specifically, the objectives of this study were : 

(i) to study the influence of species characteristics on their vulnerability to habitat loss and fragmentation, 

(ii) to assess the respective effects of current barriers to gene flow (related to deforestation and road infrastructure) and older climatic/demographic events (droughts) on the patterns of genetic diversity observed today, taking advantage of recent demographic inference methods (Approximate Bayesian computation; Beaumont et al. 2002), 

(iii) to study the effects of social structure vs. landscape fragmentation on population genetic diversity (Parreira et al. 2020 Heredity).

Roe deer project

The impact of landscape changes can vary widely between species, depending on their habitat use requirements, diet and mobility. Some native species, so-called synanthropic species, can even benefit from anthropogenic landscape changes. While many studies have investigated the effects of landscape modifications on species that are negatively impacted by habitat fragmentation, we still lack information on native species which benefit from habitat fragmentation and land use. This is, however, an important avenue of research, because many of these synanthropic species are in expansion and are the object of major societal issues (such as hunting, crop and forest damages, zoonoses, collisions with vehicules). They are thus often considered as pests by society or at least as species for which we should control their expansion and density to avoid depredation and human-wildlife conflicts. 

In addition, while the majority of studies has focused on the effects of habitat loss and fragmentation on species richness and abundance or on genetic diversity, the impacts of landscape modifications on behaviours and life-histories are still poorly understood. And yet, behavioural plasticity plays a key role in the adaptive response of species to rapid environmental changes linked to anthropogenic activities, due to the high reactivity and lability of behaviours. 

The ANR PATCH project that I coordinated during 3 years at the LBBE-CNRS therefore aimed to explain how behavioural plasticity has enabled a primarily forest-dwelling species, the European roe deer, to colonize and flourish in human-dominated landscapes such as mixed forest/farmland mosaics and open agricultural landscapes, thanks to its great adaptability and plasticity. This project was carried out in close collaboration with CEFS-INRA, CBGP-INRA, ONCFS and the Grimsö Wildlife Research Station in Sweden. More specifically, the aim was to study the behavioural, genetic and demographic response of roe deer to variations in habitat and landscape openness. 

Project name

Roe deer behavioural Plasticity and Adaptation To landscape Changes

Acronym

PATCH

Funding

Grant from the Agence Nationale pour la Recherche (ANR) - Post-Doctorant Return (RPDOC) program

(ANR-12-PDOC-0017-01)

Period

3 years and 3 months (December 2012 - March 2016)

Principal investigator

Cécile VANPÉ (LBBE-CNRS, UMR CNRS 5558)

Budget

373 514 €

Website

http://www.agence-nationale-recherche.fr/projet-anr/?tx_lwmsuivibilan_pi2[CODE]=ANR-12-PDOC-0017

This project combined field work, laboratory work and modelling and was based both on an inter-population approach (comparing 5 populations monitored over the long term in Europe with contrasting degrees of landscape fragmentation) and an intra-population approach (studying a gradient of landscape openness within a given population). 

We collated existing field data and tissue samples from the long-term monitoring (over several decades) of five contrasting roe deer populations (3 in France and 2 in Sweden), with various degree of habitat fragmentation.

We complemented these data and these samples by performing 3 new field work sessions in the three French populations in 2013, 2014 and 2015.

The field work consisted mainly in capturing roe deer, marking them individually, measuring them, collecting samples for genetic and immunological analyses and equipping animals with VHF collars to monitor their movements and survival.

We genotyped tissue samples using microsatellite molecular markers.

We used capture-mark-recapture data and population modelling in order to characterize life-history strategies and population dynamics of the different populations.

These data allowed us to characterize the behaviours, genetic diversity and demography of roe deer populations with contrasted situations of habitat fragmentation and land use.

We have highlighted the influence of body mass, genetic diversity and personality of roe deer on their decision to disperse and their dispersal distance in different habitats with various degrees of landscape openness. We have also shown that fawn survival in forest or agricultural habitats depended on the personality of their mother. We have found that the escape behaviour of deer facing danger varied as the function of the degree of landscape openness, the proximity to roads and the habitat quality. We have finally showed that the level of additive genetic variance of body mass was sensitive to environmental conditions and could strongly vary within and among populations as a function of resource availbaility.

These results show how roe deer adjust their behaviour to anthropogenic landscape changes. They will help predicting how to manage the expansion of ungulate species across Europe in the context of increasing landscape modifications.

Dispersal is a crucial mechanism enabling individuals to respond to environmental change (Ronce 2007). First of all, it provides individuals with the potential to track favourable environmental conditions. By affecting the spatial variation of genetic diversity between and within populations, it mitigates the effects of genetic drift in small populations and thus reduces the risk of extinction. Finally, gene flows induced by dispersal can in some cases promote local adaptation. In the context of global changes, it is therefore essential to better understand the major determinants (phenotypic, genetic and environmental) of variations in dispersal behaviour. We have been able to highlight the influence of body mass, neutral and adaptive (immunogenetic) genetic diversity and deer personality on their dispersal decision and distance, within different habitats with varying degrees of landscape openness (Debeffe et al. 2014 Proc R Soc Lond B, Vanpé et al. 2015 Oecologia, Vanpé et al. 2016 Oikos). We found that future dispersal individuals were less neophobic and had higher energy budgets than future philopatric individuals (Debeffe et al. 2014 Proc R Soc Lond B). Furthermore, the dispersal rate was higher in Chizé, where habitat quality is lower, than in other populations (Vanpé et al. 2015 Oecologia). The poor demographic performance and body condition of individuals in Chizé may have promoted a leave-it emergency dispersal tactic of fleeing adverse conditions. Finally, dispersal was condition-dependent only in Aurignac, where the quality of the resources was richer due to the presence of cultivated fields, and not in the forest populations (Vanpé et al. 2015 Oecologia).

An animal's flight behavior is a reliable indicator of the risk it perceives in its environment and has thus been widely used to assess how prey cope with anthropogenic disturbances. In the context of the landscape of fear, we have thus studied the behavioural plasticity of deer in their response in terms of flight behaviour to perceived spatio-temporal variations in risk (in terms of landscape openness, proximity of human infrastructure and hunting presence) and reward (in terms of habitat quality) within a fragmented agricultural landscape (Aurignac). We found that the flight behavior of deer in the face of danger varied according to the degree of openness of the landscape, the hunting season, the proximity of roads and the quality of habitat (Bonnot et al. 2017 to Anim Behav).

We studied the excursion behaviour of females during reproduction in 6 contrasting deer populations in Europe using individual GPS collar tracking (data from the collaborative project EURODEER). Within the Aurignac population, we have observed that the proportion of females making an excursion was slightly higher in the forest than in the agricultural zone, whereas the opposite pattern existed for the distance, duration and speed of excursion trajectories (Debeffe et al. 2014 Oecologia).

The personality of animals can have a significant impact on their adaptive value. Some personality types may be more adapted to particular environmental contexts. In general, we distinguish 2 main personality types related to the docility of animals: proactive individuals are more mobile and aggressive in the face of stress, while reactive individuals are more passive and react less in difficult contexts. We have shown that the fawns of proactive dams survived better in open agricultural habitats, whereas the fawns of reactive dams survived better in forest habitats (Monestier et al. 2015 Behav. Ecol.).

In the context of rapid environmental change due to human activities, the ability of individuals in a population to adapt and thrive in a new environment may depend on their ability to cope with fluctuations in their environment through the expression of individual behavioural flexibility. It is therefore important to assess the extent to which different behaviors are predictable and what factors influence their level of consistency. We have shown that docility in deer was as repeatable in the short term as in the long term and that repeatability did not differ significantly according to the age and sex of the individuals (Debeffe et al. 2015 Anim Behav). Finally, individual variation in repeatability of docility was not correlated with individual body mass.

Understanding the evolutionary and ecological mechanisms behind phenotypic responses to landscape change, and in particular whether these responses are adaptive or not, and genetically or environmentally induced, is an important challenge for evolutionary ecology. I have been closely contributing since 2015 to an innovative project of the dpt. EFPA of INRA that I co-funded with my ANR PATCH, and which is coordinated by Erwan Quéméré (CEFS-INRA), aiming at developing SNPs densely distributed in the deer genome in order to accurately measure the level of kinship between each pair of individuals as well as the inbreeding rate of populations, and to have better estimates of quantitative genetic parameters. We were thus able to study the influence of environmental conditions on the evolution of body mass in two contrasting deer populations, by reconstructing the multi-generation pedigree of these populations from microsatellite genotyping data and using a quantitative genetic approach (Animal Model) (Quéméré et al. 2018 BMC Evol Biol). We found that the level of additive genetic variance for body mass was very sensitive to environmental conditions and could vary greatly within and between populations depending on the availability of resources during early fawn growth. On the other hand, maternal genetic variance of juvenile body mass was not detectable for cohorts that experienced poor environmental conditions, but increased markedly when rearing conditions were better. The prospects are to study the genetic determinism of other morphological (antler size, body size), behavioural (personality) and life history traits (senescence, phenology) in contrasting deer populations in order to study the phenotypic response of deer to global changes.

This project was mainly driven by fundamental research, since it essentially aimed at providing new insights into how wild populations cope with environmental changes. However, the project had direct applications for society, since the results obtained will provide essential information to predict how to manage the expansion of ungulate species across Europe in relation to landscape modifications. It is a key issue from a societal, economic and environmental point of view. Indeed, while roe deer represents a major game species across Europe and hunting activity generates important revenue for many rural areas, the geographic and demographic expansion of the European roe deer has important socio-economic repercussions (i.e. forest and agricultural damage, animal-vehicles collisions, risk of expansion of human diseases vectored by deer, transmission of parasites and pathogens to cattle, impact on biodiversity). 

Brown bear project

Carnivore species are considered particularly vulnerable to the risk of local extinction in fragmented landscapes due to their relatively wide home ranges, low abundance and direct persecution by human. The sensitivity of species to landscape loss and fragmentation is heterogeneous. It depends in particular on their habitat use, diet and mobility. It is positively correlated with home range size and negatively with population density (Crooks, 2002), making small European populations of large carnivores, such as the Pyrenean Brown Bear (Ursus arctos arctos, Linnaeus, 1758), particularly sensitive to habitat loss and fragmentation.

Although in growth, the Pyrenean population of brown bear is still considered critically endangered by the IUCN, due to its low abundance, high inbreeding, fragmentation and isolation (Le Maho et al., 2013), underlining the importance of conserving and improving the functional connectivity of Pyrenean landscapes to promote genetic mixing and ensure the viability of this small population, which is also an umbrella species for many Pyrenean species and plays a key role as a keystone species in ecosystem functioning.

The aim of this study was to:

• Identify potential corridors and landscape features interesting for management and conservation of the Pyrenean brown bear,

• Examine the variations in resistance surfaces resulting from different approaches,

• Assess the differences in the linkage modelling approaches,

• Identify ways maps can be used to support decision-making.