How life-history traits affect ecosystem properties: effects of
dispersal in meta-ecosystems
Fran ç ois Massol , Florian Altermatt , Isabelle Gounand , Dominique Gravel , Mathew A. Leibold
and Nicolas Mouquet
F. Massol (http://orcid.org/0000-0002-4098-955X)(email@example.com), CNRS, Univ. de Lille, UMR 8198 Evo-Eco-Paleo, SPICI
group, FR-59000 Lille, France. – F. Altermatt and I. Gounand, Dept of Aquatic Ecology, Eawag, Swiss Federal Institute of Aquatic Science and
Technolog y, D ü bendorf, Switzerland, and: Dept of Evolutionary Biology and Environmental Studies, Univ. of Z ü rich, Z ü rich, Switzerland. –
D. Gravel, D é pt de biologie, Univ. de Sherbrooke, Sherbrooke, Canada, and: Qu é bec Center for Biodiversity Science, Quebec, Canada. – M. A.
Leibold, Dept of Integrative Biology, Univ. of Texas at Austin, Austin, TX, USA. – N. Mouquet, 7 UMR MARBEC (MARine Biodiversity,
Exploitation and Conservation), Univ. de Montpellier, Montpellier, France.
e concept of life-history traits and the study of these traits are the hallmark of population biology. Acknowledging their
variability and evolution has allowed us to understand how species adapt in response to their environment. e same traits
are also involved in how species alter ecosystems and shape their dynamics and functioning. Some theories, such as the
metabolic theory of ecology, ecological stoichiometry or pace-of-life theory, already recognize this junction, but only do
so in an implicitly non-spatial context. Meanwhile, for a decade now, it has been argued that ecosystem properties have
to be understood at a larger scale using meta-ecosystem theory because source – sink dynamics, community assembly and
ecosystem stability are all modiﬁ ed by spatial structure. Here, we argue that some ecosystem properties can be linked to a
single life-history trait, dispersal, i.e. the tendency of organisms to live, compete and reproduce away from their birth place.
By articulating recent theoretical and empirical studies linking ecosystem functioning and dynamics to species dispersal, we
aim to highlight both the known connections between life-history traits and ecosystem properties and the unknown areas,
which deserve further empirical and theoretical developments.
e study of life-history traits has primarily focused on
understanding how organism traits are aﬀ ected by the
environment and has thus used principles of evolutionary
ecology and population dynamics. is has involved basic
primary objectives such as: 1) understanding species adapta-
tions to their environment through the evolution of their life
cycle (initially dubbed as the study of life-history strategies,
Dingle 1974, Law 1979, Strathmann 1985); 2) making sense
of systematic, apparently non-adaptive phenomena such as
senescence in long-lived vertebrates or plants (Hamilton
1966, Reznick et al. 2006, Baudisch et al. 2013, Lema î tre
et al. 2015); and 3) connecting changes in organism life cycle
with their population dynamics through models of age- and
stage-structured population demographics (Charlesworth
1994, Caswell 2001). e ﬂ ip side of the issue is that life-
history traits can also be related to the eﬀ ect of organisms on
anks in part to the development of ecological theories
linking organism physiology to biogeochemical cycles, most
notably ecological stoichiometry (Sterner and Elser 2002)
and the metabolic theory of ecology (Brown et al. 2004), this
initial perspective has recently shifted to incorporate com-
plex ecological feedbacks such as ecosystem functioning and
ecological network complexity (Daufresne and Loreau 2001,
Berlow et al. 2009; Box 1 provides a glossary of concepts and
technical terms used in this paper). For example, Enquist
et al. (1999) proposed linking plant age at reproductive
maturity with biomass productivity through allometric rela-
tionships between biomass growth, standing biomass and
tissue/wood density. According to this theory, wood mass
at plant maturity should vary as the fourth power of plant
lifespan, thus allowing a rule-of-thumb to calculate the eﬀ ect
of additional extrinsic plant mortality on biomass produc-
tion. Although empirical evidence behind theories based on
allometric relationships is hard to obtain (Nee et al. 2005),
it nonetheless relates life-history traits (here, age at maturity)
with ecosystem functioning (here, plant productivity and
While ecological stoichiometry and the metabolic theory
of ecology have revealed a number of ways that life-history
can shape ecosystems (Elser et al. 2000, Berlow et al. 2009,
Hall et al. 2011, Ott et al. 2014), these hypotheses lack a
proper incorporation of ecological interactions (predation,
competition, pollination, parasitism, etc.) and do not take
the spatial structure of ecosystems into account. Other
work, most notably on host – parasite interactions and the
© 2016 e Authors. Oikos © 2016 Nordic Society Oikos
Subject Editor: Dries Bonte. Editor-in-Chief: Dustin Marshall. Accepted 5 December 2016
Oikos 000: 001–015, 2017
link between life-history strategies and organism immunity,
have succeeded in linking life-history traits to parasitic inter-
actions and ecosystem functioning through “ pace-of-life ”
syndromes (Barrett et al. 2008, R é ale et al. 2010, Wolf and
Weissing 2012, Flick et al. 2016). While pace-of-life theory-
based studies do take ecological interactions into account
to explain links between life-history traits and ecosystem
functioning, they still overlook the spatial structure of
More recently, metacommunity and meta-ecosystem
theories have improved the general understanding of the
links between the spatial structure of ecosystems and some
of their properties (Loreau et al. 2003b, Leibold et al. 2004,
Massol et al. 2011). ese include species diversity (Mouquet
and Loreau 2003, Gravel et al. 2010b), productivity
(Mouquet et al. 2002, Loreau et al. 2003a), food web inter-
actions (Amarasekare 2008), interaction network complex-
ity (Calcagno et al. 2011, Pillai et al. 2011) and stability
(Gounand et al. 2014, Gravel et al. 2016). Nevertheless,
though such theories are based on the eﬀ ects of traits on
the dynamics of communities, an explicit link between the
metacommunity literature sensu lato and life-history theories
is still lacking.
Combining metacommunity ecology with life-history
trait ecology has an obvious ‘ trait of choice ’ : dispersal i.e.
the tendency of organisms to live, compete and reproduce
away from their birth place. e aim of this article is to make
explicit the links that connect dispersal, as a life-history trait
in the population biology meaning of the word (Bonte and
Dahirel 2017), to meta-ecosystem properties using results
obtained in the ﬁ eld of metacommunity/meta-ecosystem
research. By doing so, we hope to fulﬁ l two objectives.
First we show how meta-ecosystem theory together with
other theories presented above can bridge the gap between
life-history trait studies and ecosystem properties. We then
identify remaining questions that still need to be tackled
in meta-ecosystem ecology to answer life-history driven
questions. We specify theoretical predictions that need exper-
imental testing, as well as needed theoretical developments,
to achieve an overall and coherent understanding of natural
ecosystems. Below, we ﬁ rst go through eﬀ ects of dispersal
on the functioning of meta-ecosystems. We then describe
the eﬀ ects of dispersal on the dynamics of ecosystems and
provide an empirical overview on the life-history traits driv-
ing spatial ﬂ ows between ecosystems and meta-ecosystem
properties. Finally, we conclude by discussing interactions
between dispersal and other life-history traits in the context
of meta-ecosystem ecology, and provide perspectives for
future work, both theoretical and empirical.
Dispersal and the functioning of meta-ecosystems
Ecosystem functioning is a broad class of properties that
involve ﬂ uxes and stocks of elements, energy, nutrients or
biomass among ecosystem compartments. While tradi-
tional, non-spatial ecosystem ecology considers ﬂ uxes as
the result of primary production (from abiotic compart-
ments to a biotic one), biotic interactions between species
(from a biotic compartment to another one), or death and
recycling of organic material (from a biotic compartment to
an abiotic one), meta-ecosystem ecology acknowledges the
existence of a fourth kind of ﬂ ux, i.e. ﬂ uxes due to the physi-
cal movement of biotic or abiotic material from one place to
another (Massol and Petit 2013). Because dispersal links the
Ecological stoichiometry . e study of element (e.g. carbon, nitrogen, phosphorus...) content within organisms and its
stocks and ﬂ uxes involved in ecological processes at larger scales.
Keystone and burden ecosystems . An ecosystem is said to be ‘ keystone ’ if its removal from the meta-ecosystem leads to
disproportionately deleterious consequences for a given (or several) ecosystem property (e.g. productivity) at the meta-
ecosystem scale. Conversely, a burden ecosystem ’ s removal leads to disproportionately beneﬁ cial consequences at the
meta-ecosystem scale. e deﬁ nition of ‘ disproportionately ’ in this context is based on what the removal of a typical
ecosystem of the same ‘ size ’ would entail at the meta-ecosystem scale (Mouquet et al. 2013).
Metabolic theory of ecolog y . A theory which links the diﬀ erent rates involved in organism life history (growth, consumption,
death, etc.) with body size and temperature through chemical and physical processes and laws (Brown et al. 2004).
Neutral theory of community ecology . A theory which explains the diversity of species observed in ecological communities
solely through the interplay of stochastic processes (dispersal, ecological drift, speciation, remote colonization) and not
through species niche (Hubbell 2001).
‘ Pace-of-life ’ theory of animal personality syndromes . A theory which posits that natural selection generally leads to the
existence of general personality syndromes linking physiological, immunological, foraging and life-history traits (R é ale
et al. 2010).
Resource ratio theory . A theory which explains the coexistence of species based on the complementarity of their resource
needs and their impacts on resource stocks (Le ó n and Tumpson 1975, Tilman 1980, 1982). is theory has been
expanded since then to include other limiting factors, such as predator pressure (Leibold 1995).
functioning of diﬀ erent localities, diﬀ erences in dispersal can
also change the functioning of the entire meta-ecosystem by
increasing or decreasing total primary productivity, changing
source – sink dynamics among biotic compartments or shift
the distribution of biomass across food webs (Loreau and
Initially studied as a natural extension of the insurance/
complementarity hypothesis behind the diversity-productivity
relationship (Yachi and Loreau 1999, Norberg et al. 2001),
the link between species dispersal and ecosystem productiv-
ity was ﬁ rst made explicit for a single trophic level commu-
nity in the model by Loreau et al. (2003a). e principle
behind this model is quite simple: when local environments
within patches ﬂ uctuate in time (but out-of-phase), disper-
sal allows species to average their growth rate over several
patches and, hence, to perform better than if they had not
dispersed. As explained in models of the evolution of dis-
persal in variable environments, dispersal allows ﬁ tness to
depend on its arithmetic spatial average rather than geomet-
ric temporal average (Metz et al. 1983, Massol and D é barre
2015). is better performance is immediately translated
as higher productivity when the species considered are only
primary producers (positive green arrow linking ‘ insurance ’
to ‘ primary productivity ’ through ‘ temporal variability ’ on
the right-hand side of Fig. 1).
By contrast, when the environment is spatially heteroge-
neous, but temporally constant, productivity decreases with
dispersal (Mouquet and Loreau 2003), as dispersal maintains
maladapted species through source – sink dynamics (Leibold
et al. 2004; negative green arrow linking ‘ local adaptation ’
to ‘ primary productivity ’ through ‘ quantitative spatial het-
erogeneity ’ on the right-hand side of Fig. 1). ese results
are linked to the eﬀ ects of dispersal on species coexistence:
in the absence of dispersal, local diversity is limited. At very
high dispersal, only the best species at the regional level pre-
vails. As a consequence, local diversity peaks at intermediate
dispersal, while regional diversity decreases with dispersal
(Mouquet and Loreau 2003). Both in the absence or pres-
ence of temporal ﬂ uctuations of the environment, models
based on the insurance hypothesis found positive diversity –
productivity relationships in metacommunities (Loreau et al.
2003a, Mouquet and Loreau 2003, Cloern 2007).
Primary producer coexistence, and hence productivity
following the insurance/complementarity hypothesis, might
be improved through spatial structure, i.e. the fact that
ecosystems are distinct but connected by dispersal, when
producers are constrained by more than one limiting resource
(ecological stoichiometry models; Box 2, Fig. 2). In the
models of Mouquet et al. (2006) and Marleau et al. (2015),
nutrient co-limitation, i.e. the “ perfect case ” for coexistence
in the resource-ratio theory (Tilman 1982, 1988), can be
obtained through spatial structure and dispersal only. In
such a case, resource co-limitation does not exist locally, but
emerges at a larger scale due to diﬀ erences in dispersal rates
among functional compartments (Fig. 2B). is emergent
eﬀ ect provides at the same time an explanation for increasing
primary producer growth with increasing nutrient concen-
trations in spite of potential top – down control.
It is important to consider dispersal as a life-history
trait that can diﬀ er among species within the ecosystem.
is can aﬀ ect ecosystem functioning in the same way that
heterogeneity in dispersal rates has been acknowledged,
namely as a force shaping species coexistence and diversity
distribution within ecological communities (Amarasekare
2003, Calcagno et al. 2006, Laroche et al. 2016). For
instance, Gravel et al. (2010a), found that detritus/detriti-
vore or herbivore dispersal, but not that of the basal resource,
can enhance primary productivity. Gravel et al. (2010a) also
demonstrate that the expected source – sink dynamics of
one compartment (e.g. plants) can be reversed when other
compartments (e.g. detritus or nutrients) disperse between
patches. In particular, the source – sink dynamics of primary
producers are sensitive to the balance of nutrient versus
detritus diﬀ usion; patches that would normally be unsuit-
able for them can become suitable when detritus diﬀ usion
rate is high enough (Gravel et al. 2010a; positive green arrow
linking detritus to productivity on the left-hand side of
A positive or hump-shaped relationship between dispersal
and productivity can emerge due to the dual nature of disper-
sal (i.e. as a ﬂ ux of material and energy and as a demographic
rate, Massol et al. 2011, Fig. 1). Because dispersal allows
the mixing of species across space, it tends to homogenize
composition among patches, and thus can have either a
positive or a negative eﬀ ect on productivity depending on
whether environmental variability is spatial and/or temporal
(Loreau et al. 2003a, Mouquet and Loreau 2003, see the
link between ‘ local adaptation ’ / ’ insurance ’ and ‘ characteris-
tics of limiting factors ’ in Fig. 1). By contrast, any dispersal
ﬂ ux of living organism eventually fuels the detritus pool in
the recipient patch and, hence, fertilizes it (left-hand side
arrows linking all compartments, except basal resource, to
spatial heterogeneity of limiting factors on Fig. 1). Such an
enrichment will increase regional productivity because 1) the
recipient patch becomes suitable for primary producer if it
was not in the ﬁ rst place, and 2) these ﬂ uxes make resource
use more eﬃ cient overall by preventing nutrient diﬀ usion
out of the meta-ecosystem (Gravel et al. 2010a, Fig. 1).
Other forms of interspeciﬁ c diﬀ erences may be important
in mediating spatial eﬀ ects on ecosystem functioning. For
instance, Mouquet et al. (2013) proposed the concept of
“ keystone ” and “ burden ” ecosystems, i.e. local ecosystems
that have disproportionately strong positive (for keystone)
or negative (for burden) impact on regional productivity.
Such eﬀ ects arise with spatial heterogeneity of the envi-
ronment and of nutrient inputs. Keystone ecosystems
are characterized by relatively high nutrient inputs and
dominant primary producers that have the lowest limiting
resource requirements. Because ecological stoichiometry is
likely linked to demographic parameters (Klausmeier et al.
2004), which in turn have been empirically proved to be
connected to life-history traits (Munoz et al. 2016), the
road is not long to link interspeciﬁ c variation in life-history
traits to the “ keystoneness ” of ecosystems in the framework
of Mouquet et al. (2013).
Dispersal and the dynamics of meta-ecosystems
Ecosystem dynamics refers to the temporal changes of
ecosystem variables (e.g. biomass of the diﬀ erent compart-
ments) and associated ecosystem properties (e.g. primary
productivity). At least three diﬀ erent temporal scales can
Dispersal and stoichiometry in meta-ecosystems
e resource-ratio theory of plant coexistence (Tilman 1982, 1988), based on the seminal model of Le ó n and Tump-
son (1975), has been instrumental in our understanding of the intimate linkage between stoichiometry, community
assembly and ecosystem functioning. e theory applies to two resources the R
* principle of competition theory. Its
main prediction is that stable coexistence between two species requires a particular ratio of the two limiting nutri-
ents. Owing to its accessible graphical representation, the theory has a central position in most ecological textbooks
(Begon et al. 2006). e theory has also been further developed to derive a vast array of secondary predictions, such
as the impact of resource heterogeneity and fertilization on species richness and successional dynamics (Tilman 1982,
1985). e resource-ratio theory builds on the idea that spatial heterogeneity in the ratio of limiting resources pro-
motes the maintenance of highly diverse communities (Tilman 1982). is prediction does apply to various spatial
scales, from the individual-to-individual variation in soil properties, to landscape variations. e theory does not,
however, consider the impact of spatial exchanges of plants, nutrients and other materials between localities. Both
metacommunity (Mouquet and Loreau 2002, Abrams and Wilson 2004) and meta-ecosystem (Gravel et al. 2010a)
theories in source – sinks settings have shown that the outcome of competitive interactions could be signiﬁ cantly
altered by these ﬂ ows. While it is quite challenging to elaborate a full and comprehensive theory for stoichiometry
of nutrient ﬂ ows in source – sink meta-ecosystems, it is nonetheless possible to get some intuition from a graphical
representation of two patches and two nutrients.
e graphical interpretation of the resource-ratio theory builds on a few important concepts. First, the zero
net growth isocline (ZNGI) represents the combination of the two nutrient concentrations resulting in a null
intrinsic growth rate for a given species (Fig. 2). In other words, it is the two-dimensional representation of the
* principle of competition theory. Nutrients are supplied at a given ratio in any locality, owing to processes such
as atmospheric depositions on land and river and stream inﬂ ows in lakes. In absence of consumption by primary
producers, the nutrients do equilibrate to a given concentration and ratio, represented visually as the supply point S
(Fig. 2). A key concept is that a species is able to persist provided that the supply point is located somewhere above
its ZNGI. Once a species establishes, it consumes nutrients in a given ratio, which is represented by the consump-
tion vector (the slope of the vector corresponds to the ratio of nutrient consumption). e system will converge at
equilibrium to the point corresponding to the intersection between the ZNGI and the consumption vector aligned
on the supply point. Coexistence of two species occurs provided that their ZNGIs do cross each other, and that
the supply point is located in the triangle deﬁ ned by the projection of their respective consumption vectors (Fig. 2;
Tilman 1982, 1988). When limiting factors are not resources, but natural enemies (i.e. the case in models of appar-
ent competition), the same approach can also be used (Leibold 1995, Grover and Holt 1998), although ZNGIs and
the conclusions associated with the diﬀ erent angles of intersections are not deﬁ ned in exactly the same way, and
nonlinearities in predator functional responses can lead to departures from resource-ratio theory (Grover and Holt
Nutrient cycling and any spatial exchange of nutrients between localities, whether inorganic or sequestered in bio-
mass, signiﬁ cantly complicate the situation and often make the underlying mathematics intractable. But fortunately, the
concept is pretty straightforward to illustrate graphically. In both cases, they represent an additional source of nutrient
inputs and therefore move the supply point in the two-nutrient space. In the simple case of decomposition of detritus,
where both nutrients are mineralized at the same rate, we do ﬁ nd the net supply point (S ’ ) moving away from its original
location. It increases the fertility of the system, but does not change the equilibrium situation because it keeps the same
ratio. e net supply point will, however, move in one direction or another if the mineralization or the dispersal between
localities does not respect the ratio at which it is consumed. For mineralization to alter the conditions for coexistence,
it requires that the net supply point S ’ is located within the projection of the two consumption vectors (Daufresne and
e situation is slightly more complicated for nutrient diﬀ usion, in particular when the two localities do have dif-
ferent nutrient supplies, or alternatively if they are occupied by diﬀ erent species with distinct ZNGIs and consumption
vectors. If the movement of nutrients is passive, it will move by diﬀ usion from the locality that has the highest nutrient
concentration (the source) to the locality with the lowest nutrient concentration (the sink). e location of the net sup-
ply point will therefore move in both localities. If the two localities are occupied by the same species, it will inevitably
move the supply point toward the centre of the nutrient space, as it will homogenize the meta-ecosystem. It could, how-
ever, go in the other direction depending on the characteristics of each species inhabiting localities. As a consequence,
each nutrient in a patch could thus either increase or decrease in availability, thus eventually aﬀ ecting the conditions for
an evolutionary perspective, an increase of prey extinction
rate due to predator occurrence increases the evolution-
arily stable dispersal rate in the predator, but is unimodally
linked to the evolutionarily stable dispersal rate in the prey
(Pillai et al. 2012). Overall, these results suggest that food
web assembly – and more generally ecosystem assembly
– depends on species dispersal rates in a complex fash-
ion, as predator-induced prey extinction tends to select for
more mobility in predator than in prey. When predator
presence increases prey extinction rate, foraging by the
predator can have the surprising eﬀ ect of both increasing
maximal food chain length while decreasing the average
food chain length at the metacommunity scale (Calcagno
et al. 2011).
One key ﬁ nding is that dispersal can substantially
modify theoretical predictions of ecosystem stability. May
(1972) showed with a simple model of random commu-
nity matrices that complex and diverse local ecosystems
are bound to be unstable. In contrast to May ’ s conclusion,
dispersal can substantially increase the stability of diverse
and complex ecosystems (Gravel et al. 2016). e general
principle is that dispersal tends to stabilize meta-ecosystem
dynamics because it averages responses to perturbations. As
a result, it buﬀ ers extremely strong interaction strengths,
which are the most destabilizing. e more ecosystems are
‘ spatially averaged ’ through dispersal (i.e. the more patches
are connected), the more stable the meta-ecosystem can be.
Numerical integration of Lotka – Volterra systems (Mougi
and Kondoh 2016) and individual-based simulations (Coyte
et al. 2015) lead to the same result, with the additional
eﬀ ect that very high dispersal tends to synchronize patch
dynamics and thus to ‘ homogenize ’ ecosystem responses to
perturbations, which in turn cancels the stabilizing eﬀ ect
of dispersal (Gravel et al. 2016). Hence, intermediate
dispersal rates provide the best conditions for species-rich
e eﬀ ect of dispersal on the dynamics of simple food
web modules in two-patch systems is, however, contrasted.
be distinguished. First, on long time scales, a dynamical
aspect of ecosystems is their assembly, i.e. the building-up
of ecosystems by immigration, extinction and evolution of
its component species (Morton and Law 1997). Second,
on relatively shorter time scales, the synchrony of diﬀ erent
ecosystems connected by dispersal qualiﬁ es the coherence
of diﬀ erent ecosystem dynamics (Koelle and Vandermeer
2005). Finally, on even shorter time scales, ecosystem stabil-
ity, in the sense employed by May (1972), is the tendency
of systems to return to their initial state after a small per-
turbation. ese three aspects of ecosystem dynamics are
linked in complex ways (Briggs and Hoopes 2004), and, as
we develop below, are sensitive to the amount of dispersal
Colonisation and extinction processes are at the heart
of the simplest models of ecosystem assembly. e theory
of island biogeography (MacArthur and Wilson 1963) has
been extended to food webs (Arii and Parrott 2004, Gravel
et al. 2011, Cazelles et al. 2015, Massol et al. 2017) and has
revealed rich and testable predictions (i.e. how many species,
trophic levels, etc. can be found on islands relatively to the
mainland). ese predictions arise from the interplay of two
simple rules: predators colonize islands that contain at least
one of their prey; and the extinction of a predator ’ s last prey
species entails its own extinction on an island. ese rules
result in island community assembly resembling a sampling
of the mainland food web which depends on its topology
(Arii and Parrott 2004). In the same way, the strength of
extinction cascades triggered by a single random extinction
also depends on mainland food web topology (Massol et al.
In food chains, a patch-based metacommunity model
predicts that transient food chain assembly within patches
submitted to random perturbations depends on top – down
eﬀ ects of predators on prey colonisation and extinction
rates (Calcagno et al. 2011). Longer food chains are more
likely when predator presence decreases extinction rate and
increases colonisation rate (Calcagno et al. 2011). From
Finally, organisms themselves can move across the patches. e impact of their dispersal has been extensively studied
in a wide range of conditions (Amarasekare and Nisbet 2001, Mouquet and Loreau 2002, Abrams and Wilson 2004).
Again, often the mathematics is hard to track in all of these models, but the graphical representation provides a use-
ful and general understanding of the consequences of source – sink dynamics on species coexistence. Basically, dispersal
inﬂ icts an increased loss of individuals in the location with highest density (emigration from the source), and an enrich-
ment in the location with lowest density (immigration to the sink). It provokes a translation of ZNGIs for both nutrients
(Fig. 2), moving them to higher values in the source location and to lower values in the sink. As a consequence, dispersal
might sustain a population in a location that would be otherwise inhospitable, as in traditional source – sink systems
(Pulliam 1988) or in competitive systems (Mouquet and Loreau 2002). e projection of the consumption vectors will
not be altered by dispersal of the organisms, even if there is nutrient cycling of their detritus, except in the case in which
the two nutrients are not recycled at the same rate.
In conclusion, spatial exchanges of nutrients, organisms and their detritus might alter the conditions for coexistence.
ey tend to promote regional coexistence in presence of spatial heterogeneity of supply points because 1) the supply
point moves toward the centre of the nutrient space, thereby making the conditions for coexistence more likely, and 2)
the ZNGIs move in a way that increases the tolerance of species to harsh conditions and decreases their performance in
good locations. More extensive analyses also show that it can lead to alternative stable states, and potentially dynamic
instabilities (Daufresne and Hedin 2005, Gravel et al. unpubl.). Another consequence is that dispersal, of all kinds, tends
to homogenize the meta-ecosystem in most situations.
Box 2 (Continued)
Empirical feedback to theory
Empirical work in ecology has been spurred by the theoreti-
cal development of the metapopulation and metacommu-
nity concepts, which eventually led to a better understanding
of natural ecosystems (Logue et al. 2011, Grainger and
Gilbert 2016). We are now at the point where theoretical
developments of the meta-ecosystem concept are also feed-
ing into experimental and comparative studies (Staddon
et al. 2010, Harvey et al. 2016, Gounand et al. 2017).
However, theory on meta-ecosystems is substantially more
advanced than its empirical counterpart, possibly because
of some inconsistencies between the general models and the
speciﬁ cities of natural systems (Logue et al. 2011). One such
Predator dispersal tends to synchronize and destabilize
dynamics in both predator – prey (Jansen 2001) and tri-
trophic food chains (Jansen 1995). By contrast, in nutrient –
detritus – primary producer – consumer systems, nutrient and
detritus diﬀ usion rates are destabilizing while producer and
consumer dispersal tends to be stabilizing (Gounand et al.
2014). In the latter study, intermediate consumer dispersal
rate can lead to alternative stable states of the meta-ecosystem,
with the meta-ecosystem being either in a symmetrically
oscillating state (same dynamics in the two patches) or in
an asymmetrically stable state (one patch becomes a source
of producers, consumers and detritus while the other stores
nutrients) without any underlying heterogeneity of the
environment (Gounand et al. 2014).
Effect of dispersal as
ﬂux of material/energy
Effect of dispersal as
Enrichment of scarcely
Impoverishment of scarcely
Disparate effects on limiting
factors among patches +/-
Figure 1. Links between dispersal and primary productivity according to meta-ecosystem theory (Loreau et al. 2003a, Mouquet and Loreau
2003, Gravel et al. 2010a). On the left-hand side of the diagram, dispersal of consumers, detritus and producers, seen as ﬂ uxes of material
and energy, tends to increase the amount of biomass in scarcely populated patches (i.e. those in which basal resource levels are too low for
the establishment of producers and/or consumers) and thus, through nutrient recycling, to decrease the spatial heterogeneity in nutrient
stocks among patches (blue arrow with a minus sign). Diﬀ usion of the basal resource, nutrients (Gounand et al. 2014), or producers seen
as basal resource (Pedersen et al. 2016), on the other hand, will create a source – sink movement from low-productivity patches to already
highly productive patches, thus aggravating the spatial heterogeneity of resource stocks among patches (blue arrow with a plus sign). Spa-
tially heterogeneous distribution of a single resource results in a negative eﬀ ect on primary productivity (quantitative heterogeneity). How-
ever, in case of several resources, heterogeneity in local nutrient balances (qualitative heterogeneity) may lead to positive eﬀ ects on
productivity (Marleau et al. 2015). On the right-hand side of the diagram, dispersal of primary producers seen as a demographic rate (i.e.
the I and E of the BIDE framework proposed by Pulliam 1988) generally decreases local adaptation of primary producers (they end up in
patches in which they are less well adapted, but see Edelaar and Bolnick 2012 for possible counter-examples), but primary productivity
provided by the community of primary producers gains ‘ insurance ’ against temporal variability of the environment. Dispersal thus increases
productivity at the regional scale when the environment is temporally variable, but decreases it when it is spatially heterogeneous (green
arrows going through spatial heterogeneity and spatial variability of limiting factors); the combination of the two results in a hump-shaped
link between dispersal and productivity. e blue arrows on the right-hand side of the diagram represent the potential demographic eﬀ ects
of consumer dispersal on limiting factor variability in time and space; as this eﬀ ect is quite variable across scenarios, its eﬀ ect on productiv-
ity is far from being predictable (Jansen 1995, 2001, Koelle and Vandermeer 2005, Gounand et al. 2014).
Figure 2. Graphical interpretation of the resource-ratio theory in a competitive meta-ecosystem. (A) Representation of a single ecosystem made of two resources and two competitors. Stable coexistence
will depend on where the supply point S
i is located relative to the projection of the consumption vectors C
A and C B . Equilibrium nutrient availability is indicated by the E
i and the location depends
on the ﬁ nal species composition. (B) Conceptual representation of source – sink dynamics of inorganic nutrients in a two-patch meta-ecosystem. e location of the two supply points is moved toward
the centre of the nutrient space, to locations S
, resulting in the homogenization of the metacommunity. Regional coexistence is possible in absence of nutrient movement, but not local coexistence.
e movement of net supply points toward the centre, however, allows local coexistence of the two species. (C) Representation of the eﬀ ect of dispersal of a single species on the location of the ZNGIs
in presence of source – sink dynamics. e ZNGI moves to the bottom left in the patch ii (a sink), because of immigration, allowing its stable persistence there. Similarly, the ZNGI moves to the top
right in patch i (a source), consequent to emigration.
foraging ﬁ shes (Bray et al. 1981, Schindler and Scheuerell
2002, Vanni 2002) or feces of larges herbivores or migra-
tory birds (Bazely and Jeﬀ eries 1985, Seagle 2003, Jeﬀ eries
et al. 2004), or cadavers that serve as resources in the recipi-
ent ecosystem without having a population dynamics (e.g.
migrating aquatic species in streams, Helﬁ eld and Naiman
2002, Naiman et al. 2002, Muehlbauer et al. 2014). Disper-
sal in the strict sense (Massol et al. 2011) may actually be
not feasible between diﬀ erent habitat types for most organ-
isms, as they can only live in one of these habitats and die
in the other one. In such a situation, material ﬂ ows would
be the predominant exchange. us, in many empirical sys-
tems, these material ﬂ ows are causally linked to the death
of organisms (Nakano and Murakami 2001, Sitters et al.
2015), and thus directly depend on life span as one of the
most important life-history aspects.
Most of the empirical examples of strong meta-ecosystem
dynamics involve aquatic-terrestrial linkages, in which
spatial ﬂ ows relax each other ecosystem ’ s limitations, e.g.
terrestrial carbon input into carbon-limited aquatic systems
and converse subsidy of the terrestrial system with aquatic
nitrogen (Sitters et al. 2015). A textbook example thereof
would be emerging aquatic insects, which can be accidentally
diverted into terrestrial systems during their metamorphosis
to adulthood and reproductive ﬂ ights, and subsequently die.
Importantly, these organisms, even if moving and mating
in the recipient ecosystem, oviposit in the donor ecosystem
(aquatic habitat) and do not always actively participate in
consumer-resource dynamics in the recipient ecosystem (ter-
restrial habitat) contrary to what meta-ecosystem models
assume regarding organism ﬂ ows. Flows of aquatic organ-
isms serving as resources in terrestrial systems have been
extensively described for aquatic insects but also ﬁ sh dying
after spawning (Naiman et al. 2002, Muehlbauer et al. 2014,
Sitters et al. 2015). However, these studies on strong spatial
couplings between ecosystems are mostly found in the eco-
system ecology ﬁ eld literature, with observational data either
predating or only marginally linked to the theoretical con-
cept of meta-ecosystems, which historically emerged from
the ﬁ eld of population and community ecology (Loreau
et al. 2003b).
In contrast, experimental work on meta-ecosystems has
been developed from classic experimental approaches used
for metacommunities (Logue et al. 2011, Grainger and
Gilbert 2016, Smeti et al. 2016). Such meta-ecosystem
experiments have been done almost exclusively using patches
of the same type of ecosystem (but see Venail et al. 2008 for
an example of microbial communities replicated on diﬀ erent
carbon sources), including both dispersal and mass-ﬂ ows of
resources (Howeth and Leibold 2010, Staddon et al. 2010,
Legrand et al. 2012, Livingston et al. 2012). ese experi-
ments conﬁ rm theoretical predictions that meta-ecosystem
dynamics can emerge from feedbacks between organism
dispersal and resource dynamics in same habitat-type cou-
pled systems, analogous to meta-ecosystem models (ﬁ rst
scenario in Fig. 3), such as lake or island networks, or forest
patches in an agricultural matrix. However, the important
eﬀ ects that may arise in the emblematic case studies of cou-
plings between ecosystems of diﬀ erent habitat types (second
scenario in Fig. 3) have yet to be adequately modelled or
inconsistency is the functional nature of the element mov-
ing between patches, i.e. organisms dispersing versus mate-
rial ﬂ ows. Another potential inconsistency comes from the
type of systems that are connected, because theory focuses
on ﬂ uxes among habitats of the same type, while empiri-
cists have addressed ﬂ uxes among diﬀ erent habitat types
(habitat is used in this section synonymously to the term
biotope). We here exemplify how the meta-ecosystem con-
cept is applied to empirical studies, and discuss this in the
context of life-history traits. Based on a text-book example
of meta-ecosystem dynamics, we identify possible disparities
between the theoretical work and its empirical counterparts,
and give an outlook on how to resolve the disparities and
e main focus of the metacommunity framework is
the eﬀ ect of dispersal on species coexistence, and the most
important life-history context is with respect to decisions
to disperse or not. A few dispersing individuals can often
have major consequences on the connected communities.
Implicitly, even in presence of intense habitat selection, it is
assumed that habitats are of similar kind, e.g. diﬀ erent ponds
connected by dispersal (Altermatt and Ebert 2010, Declerck
et al. 2011). is has been paralleled by extensive experimen-
tal work on metacommunities, in which same-type habitats
were connected by dispersal (Cadotte et al. 2006, Cadotte
2007, Altermatt et al. 2011, Logue et al. 2011, Grainger and
Gilbert 2016). A key ﬁ nding has been that the species traits
related to life history, such as dispersal mode or dispersal stage
induction, and life-history tradeoﬀ s can strongly aﬀ ect meta-
community dynamics and species distribution (Altermatt
and Ebert 2010, De Bie et al. 2012, Seymour et al. 2015).
ese studies, for example, found that induction of disper-
sal stages is linked to environmental deterioration inducing
speciﬁ c life-history stages (dispersal stages), and eventually
aﬀ ecting species ’ spatial distribution (Altermatt and Ebert
2010, De Bie et al. 2012). Tradeoﬀ s between competitive
ability and dispersal ability result in distributions of species
diﬀ ering from neutral models assuming otherwise identical
life-history traits (Seymour et al. 2015).
e meta-ecosystem framework explicitly considers local
nutrient dynamics and material ﬂ ows such that dispersing
organisms can also be seen as vectors of resources ﬂ owing
across units of spatial organisation. e theoretical work on
meta-ecosystems is indiﬀ erent with respect to the identity
of these habitat types. Empirically, however, there are two
major and distinct scenarios: First, the patches may be of
the same habitat type, which would then be an extension
of the metacommunity in which resource ﬂ ows would also
be added, e.g. exchange of dispersers and resources among
diﬀ erent ponds in a wetland (Howeth and Leibold 2010),
intertidal communities (Menge et al. 2015), or litter wind-
blown across diﬀ erent agroecosystems (Shen et al. 2011).
e second scenario, and possibly the most common one,
however, is that the ﬂ ows are between diﬀ erent habitat
types, such as resource ﬂ
ows between pelagic and benthic
habitats, and – more strikingly – between terrestrial and
aquatic ecosystems. Massive spatial ﬂ ows can occur between
contrasting ecosystems (Polis et al. 1997) and they are often
linked to species life-history, whereby species either trans-
port resources during foraging (e.g. seabirds on islands,
Polis and Hurd 1995) or migration, such as excretion of
traits, while in the second it arises from phenology and life
We propose that this distinction allows a better identi-
ﬁ cation of the empirical and theoretical work needed: for
same-habitat-type meta-ecosystems, we lack observational
data to adequately quantify resource ﬂ ows and we therefore
do not yet understand their signiﬁ cance for local dynamics.
For diﬀ erent-habitat type meta-ecosystems, ﬂ ows are well
documented, but theoretical models that address the role
of organisms which are not dispersing between patches but
are instead crossing the barriers to fuel recipient resource
pools are currently lacking. In pioneering modelling work,
Leroux and Loreau (2012) opened the ﬁ eld by investigat-
ing the eﬀ ects of cross-ecosystem pulsed-ﬂ ows of herbivores
as prey, but further developments in this direction are still
needed. On the experimental side, technical challenges have
to be addressed to causally separate spatial ﬂ ows of materi-
als (resources) from spatial ﬂ ows of organisms (dispersers)
(Harvey et al. 2016) in order to test precise meta-ecosystem
mechanisms. Empirical questions emerging from this sce-
nario are to test how species life-history traits in one habitat
type may cascade to other habitat types through mate-
rial ﬂ ows. Furthermore, experimental tests disentangling
interactions between perturbation regimes and spatial
ﬂ ows of resources may be highly relevant from an applied
empirical perspective, and can be addressed in an explicit
meta-ecosystem perspective. Ultimately, we expect the
dynamic interplay of theory (Loreau et al. 2003b, Massol
et al. 2011, Gounand et al. 2014) and empirical work to lead
to a more mechanistic understanding of spatial community
and ecosystems dynamics.
Dispersal: a life-history trait with many effects on
Previous sections have emphasized some ecosystem properties
that are aﬀ ected by dispersal within meta-ecosystems. First,
depending on species coexistence mechanisms, dispersal
tends to increase local diversity and meta-ecosystem pro-
ductivity, at least until intermediate levels of dispersal (Levin
1974, Mouquet and Loreau 2002, Loreau et al. 2003a,
Economo and Keitt 2008). Second, provided that patches are
suﬃ ciently heterogeneous in their response to perturbations,
dispersal stabilizes meta-ecosystem dynamics (Gravel et al.
2016, Mougi and Kondoh 2016), although the dispersal of
some trophic levels is more stabilizing than others (Gounand
et al. 2014). ird, in simple interaction networks, dispersal
tends to synchronize and destabilize local dynamics (Jansen
1995, 2001) while limited dispersal increases persistence of
otherwise ephemeral species assemblages (Briggs and Hoopes
2004). Fourth, in spatially structured heterogeneous ecosys-
tems, dispersal paves the way for nutrient co-limitation and
hence for species coexistence on a few limiting resources
(Mouquet et al. 2006, Marleau et al. 2015). On top of these
eﬀ ects of dispersal on ecosystem functioning and dynam-
ics, species dispersal/colonization abilities shape food web
complexity (Calcagno et al. 2011, Pillai et al. 2011), which
Overall, feedback of empirical observations to meta-
ecosystem theory leads to the conclusion that the drivers
of meta-ecosystem dynamics may diﬀ er depending on the
scenario of habitat types involved (Fig. 3). In same-habitat-
type meta-ecosystems, the spatial structure could be seen as
metacommunity-like, with organism dispersal as the domi-
nant spatial ﬂ ow type, and meta-ecosystem eﬀ ects would
mainly emerge from interactions between dispersal and
local resource dynamics (including local recycling). In dif-
ferent-habitat-type meta-ecosystems (e.g. aquatic-terrestrial
coupling), the spatial structure mostly consists of material
ﬂ ows (dead organisms with negligible true dispersal), and
meta-ecosystem eﬀ ects would emerge from interactions
between material ﬂ ows and local community dynamics. In
the ﬁ rst case, spatial couplings arise from species dispersal
Lakes / ponds Lakes / ponds
zoom on flowszoom on flows
LHT: phenology, life span (death)
Figure 3. Two contrasting meta-ecosystem types based on empirical
observations and illustrated by aquatic – terrestrial landscapes.
Panels (A) and (B) give examples of spatially structured landscapes
in which habitat patches are connected by spatial ﬂ ows (arrows).
Blue and green colours refer to aquatic and terrestrial respectively.
Left column shows meta-ecosystems in which patches are of same
habitat type, while right column shows meta-ecosystems in which
patches are of diﬀ erent habitat types. If we zoom on documented
ﬂ ows between two patches (bottom panels), same-habitat-type
meta-ecosystems (C) are mostly linked by organism dispersal and
potential ﬂ ows of resource (R), but these are poorly documented
(dotted arrows). Diﬀ erent-habitat-type meta-ecosystems (D) are
linked by exchanges of dead organisms fuelling the resource pool.
Aq , and B
T refer to biomass of aquatic and terrestrial organisms
tradeoﬀ preventing an organism from both moving and
eating at the same time). A second possibility is that dis-
persal correlates with the other trait because both traits are
structurally linked, e.g. they both scale with organism size
(allometric link) or they both respond similarly to biological
stoichiometric changes (stoichiometric link). Finally, both
dispersal and the other trait can be shaped by joint selective
pressures, with either the same pressures acting on both traits
at once (e.g. dormancy and dispersal, Vitalis et al. 2013) or
one or both trait(s) having a selective feedback on the other
(e.g. selﬁ ng and dispersal, Cheptou and Massol 2009, or local
adaptation and dispersal, Berdahl et al. 2015). In practice,
correlations between dispersal and other life-history traits
can only be uncovered when there is suﬃ cient variation in
the traits under study, which means that the wider ‘ the phy-
logenetic net ’ , the easier it is to capture such correlations.
However, interpreting these correlations as resulting from
tradeoﬀ s, structural constraints or joint evolution is often
diﬃ cult and experimentally challenging, especially when the
problem is framed as the inference of life-history invariants
(Nee et al. 2005).
It would be diﬃ cult to enumerate here all the possibilities
of dispersal-trait correlations that would likely have impacts
on meta-ecosystem functioning and dynamics. Some of
these have already been considered separately. For example,
Otto et al. (2007) ’ s study on the eﬀ ect of predator – prey body
mass ratios on food web stability could be easily coupled
with Gravel et al. ’ s (2016) study on the eﬀ ect of dispersal on
ecosystem stability to gain insight into the combined eﬀ ects
of dispersal and body size when both traits are structurally
linked. Others readily lend themselves to speculation. For
example, with higher passive dispersal in smaller organisms
and the relationship between initial growth, asymptotic size
and temperature in ectotherms (Atkinson et al. 2006), one
is tempted to think that warming oceans might become
less connected by dispersal, as some data on larval dispersal
already suggest (O’Connor et al. 2007), which in turn would
aﬀ ect their functioning and dynamics as predicted by the
models described in previous sections.
An especially challenging issue regarding life-history trait
evolution and meta-ecosystem properties is to link ecologi-
cal stoichiometry with ecosystem properties through cell and
organism physiology (Jeyasingh and Weider 2007), e.g. as
proteins and rRNA have diﬀ erent stoichiometry (Loladze
and Elser 2011). For instance, the proportion of phosphorus
content due to RNA (versus due to skeleton) is expected to
decrease with body mass in vertebrates (Gillooly et al. 2005).
In some insects, high-dispersal genotypes are associated
with particular alleles at genes coding for phosphoglucose
isomerase (PGI), e.g. in the Glanville fritillary butterﬂ y
(Haag et al. 2005, Hanski and Saccheri 2006). Eﬃ cient PGI
genotypes have a higher peak metabolic rate and ﬂ y longer
than less eﬃ cient types (Niitep õ ld et al. 2009, Niitep õ ld and
Hanski 2013). As the PGI enzyme is involved in glycoly-
sis and gluconeogenesis, a link between PGI and ecological
stoichiometry might be expected (as suggested by experimen-
tal evidence on Daphnia pulex , Jeyasingh and Weider 2005,
Weider et al. 2005) which, in turn, would link ecological
stoichiometry with dispersal ability. is ﬁ eld of inquiry is
just beginning, but might reveal exceptional ﬁ ndings linking
traits and ecosystem functioning, such as an increased spatial
can potentially feedback on ecosystem stability (Allesina and
Tang 2012, Neutel and orne 2014, Grilli et al. 2016).
Meta-ecosystem theory is not solely geared towards
understanding the functioning of ecosystems, but also
grounded in the foundations laid out by metapopulation
and metacommunity theories. erefore, the movements
of species within a meta-ecosystem are bound to be gov-
erned by how organisms perceive their environment and
where they thrive – i.e. non-random dispersal, habitat selec-
tion, foraging and dispersal evolution (Amarasekare 2008).
e feedback of meta-ecosystem state on dispersal evolu-
tion has just begun to be studied, and has focused so far
on simple predator – prey conﬁ gurations (Chaianunporn and
Hovestadt 2012, Pillai et al. 2012, Drown et al. 2013, Travis
et al. 2013, Amarasekare 2015). On top of all the mecha-
nisms of dispersal evolution that are already known (Bowler
and Benton 2005, Ronce 2007, Duputi é and Massol 2013),
meta-ecosystem context is likely to provide new selection
mechanisms through the discrepancy in generation time
and spatial scale of motility of diﬀ erent trophic levels. For
instance, dispersal is selected against when environmen-
tal quality of habitat patches is positively autocorrelated in
time, but selected for when it is positively autocorrelated in
space (Travis 2001, Massol and D é barre 2015). In the case
of a prey species for which predator presence is an ‘ environ-
mental characteristic ’ , as predators live longer, have slower
population dynamics and can cover and forage over sev-
eral prey patches at once, the ‘ eﬀ ective ’ autocorrelation of
the environment for the prey will likely be positive in both
time and space, thus aﬀ ecting the evolution of prey dispersal.
Evolution of dispersal in food webs also imposes a feedback
between the cost of dispersal and dispersal itself, as sparse
prey populations can diminish predation pressure and,
hence, decrease the cost of dispersal borne out of predation
between habitat patches. Finally, it is also noteworthy that,
even though dispersal evolution has begun being considered
in a food web context, the consequences of this evolution on
ecosystem functioning have yet to be studied.
Other life-history traits and their impact on
e central tenet of meta-ecosystem studies is that species
dispersal may be responsible for many observations that
would otherwise require more complicated theories to
explain, such as the maintenance of maladapted species (the
“ mass eﬀ ect ” paradigm of metacommunity theory, Shmida
and Wilson 1985, Leibold et al. 2004) or the distribution
of species abundance in ecological samples, as predicted
by the neutral theory of ecology (Hubbell 2001, Volkov
et al. 2003). From this central tenet, it is no wonder that
the main connection made by these studies between life-
history traits and ecosystem properties considers dispersal as
the life-history trait of interest. However, dispersal generally
correlates with a wide palette of other traits (e.g. fecundity,
body size, etc., Bonte and Dahirel 2017), known collectively
as “ dispersal syndromes ” (Clobert et al. 2009, Ronce and
Clobert 2012, Duputi é and Massol 2013). Such correla-
tions can be explained in three ways: ﬁ rst, the other trait
can correlate with dispersal ability because there is a tradeoﬀ
constraining the values of both traits (e.g. time allocation
Experimental studies are required to explore whether 2)
diﬀ erent ecosystem functions are aﬀ ected diﬀ erently by
the movements of nutrients, detritus, primary produc-
ers, consumers, etc. Existing models suggest that disper-
sal asymmetries can do more than just alter source-sink
dynamics (Gounand et al. 2014) and existing experiments
point out possible eﬀ ects of basal species dispersal on spe-
cies regulation processes (Howeth and Leibold 2008).
e general prediction that intermediate dispersal rates 3)
should stabilize meta-ecosystems has to be tested prop-
erly, both experimentally (but see Howeth and Leibold
2010, 2013), and based on large-scale observational data-
sets of abundance time series (following the approach of
Jacquet et al. 2016).
e idea of spatial complementarity between habitats 4)
within a meta-ecosystem needs to be assessed and experi-
mentally challenged. For instance, when ecosystems are
intrinsically limited by diﬀ erent nutrients in diﬀ erent
habitats (e.g. C in aquatic habitats versus N in terrestrial
ones), experiments are needed to assess whether interme-
diate (or high) spatial ﬂ ows of biotic compartments lead
to higher productivity.
Experiments should test whether spatial structure and 5)
heterogeneity of supply points can lead to the stable
coexistence of species with diﬀ erent resource ratios
(Mouquet et al. 2006, Marleau et al. 2015), possibly
exploring situations more complicated than two-patch,
two-species, two-resource systems.
eoretical studies are needed to explore how perturba-6)
tions propagate within a meta-ecosystem, depending on
which compartments are dispersing more, on connectiv-
ity patterns, on ﬁ rst-disturbed compartments and on the
nature of the perturbation (invasion, extinction, habitat
destruction, etc.), following new perspectives on the
notion of stability in ecology (Arnoldi et al. 2016).
One promising theoretical endeavour would be to pre-7)
dict the impact of ecosystem removals on diversity and
functioning in a spatially explicit fashion, thus merging
models of Economo and Keitt (2008, 2010) on diversity
in metacommunity networks and Mouquet et al. (2013)
on keystone ecosystems.
Species coevolution models are highly needed to assess 8)
whether evolution leads to increases or decreases in
productivity, ﬂ uxes, synchronicity, stability, etc. at the
meta-ecosystem scale, e.g. focusing on the evolution of
dispersal at diﬀ erent trophic levels within food webs.
Models of ecosystem assembly and disassembly should 9)
be developed to assess the conditions of existence of
“ forks ” (i.e. alternative trajectories), “ dead-ends ” or
loops in the topology of ecosystem successions (Law and
Acknowledgements – We thank D. Bonte for organizing this special
issue and allowing FM to present the idea of the paper at the Nordic
Oikos Society meeting in Turku, February 2016. We thank Dries
Bonte for insightful comments on an earlier version of the
Funding – FM was supported by the CNRS and through the
ANR-funded project ARSENIC (ANR-14-CE02-0012). DG was
supported by the NSERC and the Canada Research Chair Program.
diﬀ usion of one type of nutrient over another one due to a
systematic association of body stoichiometry with dispersal
Challenges ahead for meta-ecosystem ecology
e study – both theoretical and empirical – of mecha-
nisms linking organism dispersal and ecosystem properties
is a recent endeavour in ecology. To date, meta-ecosystem
ecology has focused on linking community ecology (spe-
cies coexistence, distribution of diversity), with ecological
dynamics and demographics (ecosystem stability, synchrony,
assembly), ecological interaction networks (network com-
plexity, material/energy ﬂ uxes) and functional ecology
(stocks, ﬂ uxes and productivity). However, two interfaces
have yet to be strengthened with respect to life-history traits
and meta-ecosystem properties.
First, the integration of biogeography and functional
ecology through meta-ecosystems has only begun to be
addressed (Wieters et al. 2008, Meynard et al. 2011, Kissling
et al. 2012, Nogales et al. 2016). is interface between
meta-ecosystem ecology and biogeography is a necessary
step if we are to extend species distribution models and other
map-based representations of biodiversity to map-based rep-
resentations of ecosystem functioning and link these with
the underlying mechanisms involved. As life-history traits
play key roles in determining species response to anthropi-
cally driven changes of the environment (Lindborg 2007,
Colautti et al. 2010, Ojanen et al. 2013), life-history traits,
and dispersal in particular, will probably play a key role in
explaining spatial distribution of ecosystem functioning.
Second, we can ask whether variability in life-history
traits such as dispersal may entail direct consequences for
ecosystem properties. For instance, Laroche et al. (2016)
recently studied the evolution of dispersal in a model based
on Hubbell ’ s (2001) neutral model of biodiversity to assess
whether species would converge or diverge in dispersal rate.
As it turned out, diversity patterns are strongly altered by
disruptive selection on dispersal (Laroche et al. 2016). Spec-
ulation linking these results with others from meta-ecosystem
models (Gounand et al. 2014) may lead us to think heteroge-
neous selection on dispersal rates among trophic levels could
drive eco-evolutionary feedbacks linking dispersal evolution
and ecosystem functioning.
Closing words: empirical and theoretical directions
We list here several important directions that deserve further
enquiry, both on the empirical and theoretical fronts. Meta-
ecosystem ecology and its interface with life-history studies
in particular need to be strengthened by making experi-
ments to test important meta-ecosystem predictions and by
developing meta-ecosystem models in directions that will
more strongly link them to life-history traits:
e maximization of ecosystem productivity at inter-1)
mediate dispersal has to be tested with respect to the
mechanisms maintaining coexistence of primary produc-
ers (Mouquet et al. 2002, but see Howeth and Leibold
2008), and the eﬀ ect of dispersal asymmetries between
trophic levels on productivity (Gravel et al. 2010a) needs
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MAL was supported by NSF-DEB 1353919. NM was supported
by the CNRS. FA and IG were supported by Swiss National Sci-
ence Foundation Grant PP00P3_150698, University of Zurich and
is work was inspired by early discussions between FM, DG
and MAL at the ‘ Stoichiometry in meta-ecosystems ’ working group
at the National Institute for Mathematical and Biological Synthesis,
sponsored by the National Science Foundation, the US Dept of
Homeland Security, and the US Dept of Agriculture through NSF
Award no. EF-0832858, with additional support from e Univ. of
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