Download Predator control of ecosystem nutrient dynamics

Ecology Letters, (2010) 13: 1199–1209
IDEA AND
PERSPECTIVE
Oswald J. Schmitz,1* Dror
Hawlena1 and Geoffrey C.
Trussell2
1
School of Forestry and
Environmental Studies,
Yale University, New Haven,
CT 06511, USA
2
Marine Science Center,
Northeastern University,
Nahant, MA 01908, USA
*Correspondence: E-mail:
[email protected]
doi: 10.1111/j.1461-0248.2010.01511.x
Predator control of ecosystem nutrient dynamics
Abstract
Predators are predominantly valued for their ability to control prey, as indicators of high
levels of biodiversity and as tourism attractions. This view, however, is incomplete
because it does not acknowledge that predators may play a significant role in the delivery
of critical life-support services such as ecosystem nutrient cycling. New research is
beginning to show that predator effects on nutrient cycling are ubiquitous. These effects
emerge from direct nutrient excretion, egestion or translocation within and across
ecosystem boundaries after prey consumption, and from indirect effects mediated by
predator interactions with prey. Depending on their behavioural ecology, predators can
create heterogeneous or homogeneous nutrient distributions across natural landscapes.
Because predator species are disproportionately vulnerable to elimination from
ecosystems, we stand to lose much more from their disappearance than their simple
charismatic attractiveness.
Keywords
Consumptive effects, ecosystem function, ecosystem services, indirect predator effects,
non-consumptive effects, nutrient cycling, nutrient translocation, predator behaviour and
nutrient distribution.
Ecology Letters (2010) 13: 1199–1209
INTRODUCTION
Species losses within ecosystems tend to be heavily biased
towards predators (Pauly et al. 1998; Duffy 2003; Myers &
Worm 2003; Halpern et al. 2005; Dobson et al. 2006;
Heithaus et al. 2008). Predators often exert a strong
influence on ecological communities by controlling the
abundance and dynamics of species in lower trophic levels
(Soule et al. 2005; Wootton & Emmerson 2005; Dobson
et al. 2006; Thebault & Loreau 2006). Together, the
susceptibility of predators to loss and the pivotal role they
often play in natural systems will likely jeopardize the level
of overall services that they provide.
Predators tend to be valued for the provision of three
broad services: their ability to regulate prey populations
(Halpern et al. 2005; Dobson et al. 2006), as indicators of
high species richness that warrant conservation (Sergio et al.
2006) and, given their charismatic nature, their economic
contribution to recreation and ecotourism (Dobson et al.
2006). Unfortunately, this valuation of predators treats their
losses merely as early indicators of environmental destruction that may lead to the eventual loss of species in lower
trophic levels. Currently, it is believed that species in lowest
trophic levels are most essential to critical life-support
services in ecosystems (e.g. ecosystem production, nutrient
cycling, carbon storage; Dobson et al. 2006).
This line of reasoning implies weak, if any, connections
between the loss of top predators and the loss of critical lifesupport services like nutrient cycling. However, this view
requires considerable revision (Polis et al. 1997, 2004; Vanni
2002). Mounting research reveals that top predators can
have cascading effects that extend beyond their prey base to
impact on ecosystem nutrient dynamics that may then feed
back upward to influence the biological productivity.
Ignoring this connection may mean that an important
functional role of predators in ecosystems remains grossly
undervalued, especially given that nutrient cycling is deemed
to be among the most valuable of all ecosystem services
(Costanza et al. 1997).
BOTTOM-UP VS. TOP-DOWN CONTROL
OF NUTRIENT DYNAMICS
The classical view of ecosystem functioning holds that
microbial species are the most critical driver of nutrient
dynamics owing to their capacity to convert organic matter
into mineral elements for plant uptake and production. This
view gives primacy to bottom-up control of nutrient
2010 Blackwell Publishing Ltd/CNRS
1200 O. J. Schmitz, D. Hawlena and G. C. Trussell
dynamics, and ecosystem functioning more generally,
because microbial action is believed to be the rate-limiting
step in the delivery of inorganic nutrients for primary
and secondary production in nutrient-limited systems
(Schlesinger 1991). However, animals can directly and
indirectly control the fate of nutrients in ecosystems, and
may sometimes circumvent the need for microbial mineralization altogether, depending on how and where nutrients
are consumed, translocated and eliminated (Vanni 2002;
Wardle & Bardgett 2004). But, the relative strength of
animal vs. microbial control over nutrient cycling can
depend on environmental context. For example, nutrient
cycling may be more strongly affected by microbes in
benthic habitats whereas animals may play a more important
role in pelagic environments.
Animals exert direct control whenever consumed
nutrients are either assimilated into their body tissue or
egested as faeces (Vanni 2002; Wardle & Bardgett 2004).
Assimilated nutrients in turn are allocated to secondary
production (growth and reproduction), excreted via urine
or an equivalent waste elimination system (e.g. in guano),
or in the case of C released via respiration (Kitchell et al.
1979; Vanni 2002). Carcasses that become part of the
detrital pool also eventually release nutrients that are
bound up in animal body tissue. Animals can also
indirectly influence the fate of nutrients through selective
feeding on resources. Selective feeding, driven by physiological demands to maintain specific body elemental
ratios of C, N and P, can lead to changes in the elemental
content of resource species (Kitchell et al. 1979; Carpenter
et al. 1992; Sterner & Elser 2002; Vanni 2002; Wardle &
Bardgett 2004; Schmitz 2010). Such selective feeding alters
the nutrient content (quality) of resource tissues that
eventually are released as detritus to the dead organic
matter pool.
These different forms of released nutrients mean that
animals alter the fate of nutrient supplies by varying the
availability and thus the propensity of nutrients to enter the
ÔfastÕ vs. ÔslowÕ cycle (McNaughton et al. 1988; Ritchie et al.
1998; Vanni 2002; Shurin et al. 2006; Schmitz 2008).
Nutrients enter the fast cycle whenever animals excrete
them in inorganic forms that can be readily taken up by
autotrophs. Slow cycling occurs when animals egest
nutrients in organic form or alter the tissue elemental
composition (i.e. quality) of plant and animal matter that
must then be decomposed and mineralized before nutrients
can be taken up by autotrophs. Effects of fast cycling are
often evident within a single growing season (McNaughton
et al. 1988; Vanni 2002). Slow cycling tends to become
evident after one or several seasons (McNaughton et al.
1988; Vanni 2002) but recent evidence suggests that it may
also become evident within a single season (Spivak et al.
2009).
2010 Blackwell Publishing Ltd/CNRS
Idea and Perspective
The influence of predators
Early evidence that predators may regulate nutrient cycling
emerged largely from studies in lakes that discovered four
broad mechanisms of top predator effect (Carpenter et al.
1992; Vanni 2002): (1) predation that leads to spatial and
temporal shifts in the size and identity of living zooplankton
grazers and hence nutrients contained in the zooplankton
community. This arises either because C : N : P varies with
prey species or because size-selective predation shifts preysize distributions and hence nutrient cycling rates because
cycling rates scale allometrically with prey size. (2) Trophic
cascade effects on phytoplankton size structure due to the
size-selective predation on zooplankton that leads to sizedependent effects on rates of nutrient turnover through the
phytoplankton. (3) Excretion and egestion of nutrients by
predators directly into the water column (habitat); and
(4) the spatial translocation of nutrients via consumption
of resources in one location and excretion or egestion into
another.
The most comprehensive syntheses, to date, of animal
control of nutrient dynamics have covered freshwater
(Vanni 2002) and terrestrial (Wardle & Bardgett 2004)
ecosystems, but evidence of predator effects on nutrient
cycling in these reviews is only addressed for freshwater
systems (Vanni 2002). Substantial evidence for predator
effects in a host of other ecosystems has accumulated
since these reviews were published. Using this emerging
published literature, we offer an evidence-based elaboration of the earlier idea (Kitchell et al. 1979) that predator
effects on nutrient dynamics should occur in all ecosystems. We identified the literature first using Web of
Science with various concatenations of the key words
predator, ecosystem and nutrients. We identified additional
literature from reference sections of the Web of Science
sources. Here, we included only those studies that were
not included in the synthesis by Vanni (2002). Furthermore, we did not include studies that model effects, but
instead only present evidence from studies that empirically
demonstrated a link between predators and the fate of
nutrients.
Our search revealed that predators may control nutrients
via fast and slow cycling in many ecosystems through the
classic mechanisms described above as well as others that
are now only becoming evident after deliberate efforts to
study trophic interactions and ecosystem function (Table 1).
Many of these mechanisms have not yet been considered
by current theory (e.g. DeAnglis 1992; Polis et al. 1997;
Leroux & Loreau 2008). We describe how these mechanisms
operate and identify unanswered questions and issues in
order to encourage more research that deliberately integrates
trophic chain interactions – the domain of community
ecology – with biogeochemical cycling – the domain of
Idea and Perspective
Table 1 Mechanisms of top predator effect on ecosystem nutrient
dynamics
Consumptive effects
(1) Alteration of C : N : P content of the prey community via
size-selective predation
(2) Trophic cascades along the plant- or detritus-based chain
(3) Consumption and release of nutrients by predators within the
same habitat
(4) Translocation of consumed nutrients across habitat boundaries
(5) Decoupling carcass distribution from live-prey distribution
(6) Alteration of prey nutrient transport and release via prey
capture
Non-consumptive effects
(7) Spatial and temporal redistribution of nutrients via predatorinduced changes in prey habitat shift
(8) Alteration of community composition and nutrient dynamics
via predator-induced herbivore foraging shifts
ecosystem ecology (Carpenter & Kitchell 1988; Vanni 2002;
Schmitz 2008).
MECHANISMS OF PREDATOR EFFECT
ON NUTRIENT DYNAMICS
Predator effects on ecosystems can emerge via consumptive
or non-consumptive pathways (Schmitz 2010). Consumptive
effects arise when predators consume nutrients within prey
and physically store, translocate and release them to the
environment. Non-consumptive effects arise when predators
elicit antipredator responses in prey that are manifest in
either or all of three general ways: habitat shifts that provide
refuge from predators, diet shifts that balance trade-offs
between foraging and risk avoidance, and stress-induced
changes in metabolism that change the demand for and
release of particular nutrients. Predator consumptive effects
are readily observable in field studies, whereas non-consumptive effects tend to be more subtle and thus require
systematic predator-exclusion experiments to be revealed.
We identified four new mechanisms (mechanisms 5–8;
Table 1) that, when combined with the original four
(mechanisms 1–4; Table 1), lead to eight different ways
that predators may control nutrient dynamics through
consumptive and non-consumptive means. In the following,
we highlight the ecology underlying the mechanisms. To do
so, we aggregated publications that arose from the same
study system – typically they provided complementary
evidence – in order to maintain independence among the
case examples (summarized in Table 2). We note at the
outset that our search did not reveal new evidence for
size-selective predation effects (mechanism 1; Table 1) in
non-pelagic systems. However, there was new evidence for
the remaining seven mechanisms.
Predator control of ecosystem nutrient dynamics 1201
Classic mechanisms
In the following sections, we add to the well-synthesized
(Vanni 2002) cases of predator effects in freshwater systems
by updating with additional freshwater examples or expanding with examples from other ecosystems.
Trophic cascades along the plant- or detritus-based chain
We did not find new evidence that predators influence
nutrient cycling via classic consumptive effects propagating
along the predator–herbivore–plant–organic matter pool
chain. However, there was emerging evidence that predator effects may influence nutrient dynamics via cascading
effects along the detritus-based chain where predation on
detritivores alters decomposition and mineralization rates
(Schmitz 2010). Although some have forcefully argued that
the detrital chain is almost entirely bottom-up controlled
(Mikola & Setala 1998; Moore et al. 2004), meta-analysis of
trophic interaction strength revealed that trophic cascades
occurred in 50% of the cases examined by Schmitz (2010).
In three-level chains comprised of arthropod predators and
prey, top predators caused a 1.2- to 3-fold reduction in
decomposition and mineralization rate, whereas in four-level
chains predators enhanced decomposition by 20% (Schmitz
2010). An exclusion experiment in a tropical forest found
that vertebrate predators enhanced soil inorganic P by
1.2- to 1.5-fold in four-level chains comprised of top
predators (birds and mammals), spider carnivores, arthropod microbivores and microbes decomposing organic
matter (Dunham 2008).
Consumption and release of nutrients by predators
within the same habitat
Persson & Svensson (2006) showed that benthivorous fish
can cause a 1.5- to 1.8-fold increase in inorganic P and
1.3- to 1.5-fold increase in inorganic N concentration in
the water column of lakes. Inorganic nutrient release from
predators has also been detected in terrestrial systems. For
example, insectivorous frogs (Eleutherodactylus coqui) release
ammonium N and P on vegetation. This release rapidly
increases nutrient concentrations on the exterior of forest
leaves and in litter by 1.4- to 2-fold over conditions where
these predators are absent and, in turn, indirectly enhances
decomposition rates (Sin et al. 2008).
Translocation of consumed nutrients across habitat boundaries
Top predators often range widely between foraging bouts
or move seasonally among geographic locations. Consequently, they have much potential to disperse nutrients
widely across habitats within ecosystems (Vanni 2002)
and across ecosystem boundaries (Polis et al. 1997,
2004). However, the temporal and spatial distribution of
nutrient inputs can depend on the behavioural ecology of
2010 Blackwell Publishing Ltd/CNRS
1202 O. J. Schmitz, D. Hawlena and G. C. Trussell
Idea and Perspective
Table 2 Summary of studies demonstrating predator effects on nutrient cycling
Predator
Ecosystem
Trophic cascades along the plant- or detritus-based chain
Insectivorous birds and mammals
Terrestrial–tropical forest
Biophysical processes
Ref.
Lower inorganic P
1
Consumption and release of nutrients by predators within the same habitat
Fish-Tench (Tinca tinca)
Aquatic-pond
Increased SOM; increased porosity;
increased N uptake in the sediment;
Bream (Abramis brama)
increased mineral N concentration in water
column; decreased NH4+ concentration in
water; higher P concentration in water
Invasive frog
Terrestrial–Oceanic island
Increased N and P concentration in leaf
(Eleutherodactylus coqui)
washes; increased Mg, N, P, K in
decomposing leaf litter
Translocation of consumed nutrients across habitat boundaries
Sea birds
Marine and terrestrial oceanic islands
Sea birds
Marine and terrestrial oceanic islands
Sea birds
Marine and terrestrial oceanic islands
Great cormorant
(Phalacrocorax carbo)
Great cormorant
(Phalacrocorax carbo)
Riparian
Crow (Corvus corone &
Corvus macrorhynchos)
Brown bears (Ursus arctos)
Haemulid fishes
Residential area to forest
Alewives (Alosa oseudoharengus)
Marine and freshwater streams
Salmon (Oncorhynchus spp.)
Marine and freshwater streams
Loggerhead sea turtles
(Caretta caretta)
River otters (Lontra canadensis)
Loggerhead sea turtles
(Caretta caretta)
Marine and terrestrial dune
Temperate forest
Marine and terrestrial riparian
Marine pelagic and coral reef
River to riparian
Marine and terrestrial dune
Decoupling carcass distribution from live-prey distribution
Brown bear (Ursus arctos)
Aquatic and riparian
Gray wolf (Canis lupus)
Terrestrial forest
Brown bear (Ursus arctos)
Black bear (Ursus americanus)
Limpkin (Aramus guarauna)
3
Increased marine-derived N
Increased soil P; increased foliar N
Increase total C, N, P; increase available C,
N; slower litter decomposition
Increased total P; increased plant available P
4
5, 6
7
Increased forest floor and mineral soil P;
increased mineral soil N; decreased litter
decomposition
Increased N, P
9, 10
Increased N loading to riparian forest
Increased N, P loading from pelagic fish to
coral reef
Increased marine N, P loadings into
freshwater streams
Increased NH4+; increased N, P, Ca
concentration
Increased soil organic matter N, P from eggs
to beach
Increased foliar N in latrine sites
Increased soil and foliar N
13
14
8
11, 12
15
16, 17
18
19
20
21
22
Aquatic–terrestrial (riparian)
Increased soil N; increased N2O flux
Increased inorganic N, P, K; increased foliar
N; decreased foliar C : N ratio
Increased soil N
Aquatic wetlands
Increased P; increased foliar N, P
24
Decrease total C, N, P; decreased available
C, N; higher litter decomposition
Lower soil P; lower foliar N
Enhanced C sequestration in live plants;
reduced C sequestration in non-living
pools; increased total C storage; reduced N
concentration of foliage and litter; reduced
release of N from decomposing litter;
lower litter decomposition
7
Alteration of prey nutrient transport and release via prey capture
Rats (Rattus rattus and Rattus
Terrestrial–oceanic islands
norvegicus)
Arctic fox (Alopex lagopus)
Terrestrial–oceanic islands
Rats (Rattus spp.)
Terrestrial–oceanic islands
2010 Blackwell Publishing Ltd/CNRS
2
23
5, 6
25, 26
Idea and Perspective
Predator control of ecosystem nutrient dynamics 1203
Table 2 continued
Predator
Ecosystem
Biophysical processes
Ref.
Damselfly larva
(Mecistogaster modesta)
Sheet-web spinning spiders
(Linyphiidae)
Aquatic–Bromeliad
Increased flow of N from litter to
Bromeliads
Increased N, P, K input into the habitat
27
Terrestrial–glacial moraines
Spatial and temporal redistribution of nutrients via predator-induced changes in prey habitat shift
Gray wolf (Canis lupus)
Terrestrial (Yellowstone)
Decreased N mineralization
Fish (Rutilus rutilus)
Aquatic–benthic
Increased rate of mineralization (Aerating
the sediment; feeding on detritus
redistributing particles)
Alteration of plant community composition and nutrient dynamics via predator-induced herbivore foraging shifts
Spider (Pisaurina mira)
Terrestrial
Increased N mineralization rate
Blue crab (Callinectes sapidus)
Aquatic–benthic
Shifts in fatty acid and labile C composition
of sediment organic matter
28
29
30
31
32, 33, 34
Studies are grouped by mechanism of predator effect.
(1) Dunham (2008); (2) Persson & Svensson (2006); (3) Sin et al. (2008); (4) Barrett et al. (2005); (5) Croll et al. (2005); (6) Maron et al. (2006);
(7) Fukami et al. (2006); (8) Breuning-Madsen et al. (2008); (9) Hobara et al. (2005); (10) Osono et al. (2006); (11) Fujita & Koike
(2009); (12) Fujita & Koike (2009); (13) Hilderbrand et al. (1999); (14) Meyer & Schultz (1985); (15) Post & Walters (2009); (16) Gende et al.
(2002); (17) Naiman et al. (2002); (18) Bouchard & Bjorndal (2000). (19) Crait & Ben-David (2007); (20) Hannan et al. (2007); (21) Holtgrieve
et al. (2009); (22) Bump et al. (2009); (23) Gende et al. (2007); (24) Macek et al. (2009); (25) Wardle et al. (2007); (26) Wardle et al. (2009); (27)
Ngai & Srivastava (2006); (28) Hodkinson et al. (2001); (29) Frank (2008); (30) Stief & Holker (2006); (31) Schmitz (2006); (32) Canuel et al.
(2007); (33) Spivak et al. (2007); (34) Spivak et al. (2009).
predators, such as whether they live solitarily, in small
groups or packs vs. in colonies or large migratory
aggregations.
Most studies have examined nutrient translocation from a
donor to an adjacent, recipient ecosystem. In these cases,
solitary or small groups of predators tend to increase
heterogeneity in local nutrient supply rates (i.e. local
nutrient hotspots) and hence heterogeneity in primary
production within the recipient ecosystem. Grizzly bears
(Ursus arctos) and river otters (Lontra canadesnis) consume a
variety of fish species within rivers. The fish either originate
in situ in these rivers (Crait & Ben-David 2007) or they
migrate to the rivers from the marine realm (Hilderbrand
et al. 1999; Crait & Ben-David 2007; Gende et al. 2007;
Holtgrieve et al. 2009). These predators release freshwater
or marine-derived nutrients (i.e. N or P) up to 1000 m
beyond the riparian zone into localized latrine (inorganic
form) or defecation (organic form) sites (Hilderbrand et al.
1999; Crait & Ben-David 2007; Gende et al. 2007;
Holtgrieve et al. 2009). This localized contribution to both
the fast- and slow-cycle creates rapid and sustained
heterogeneity in soil and foliar N and P concentrations as
well as primary production across the landscape. For
example, mean redistribution rates of salmon-derived N by
adult female brown bears were 37.2 ± 2.9 kg year)1 per
bear (range 23.1–56.3), of which 96% (35.7 ± 2.7 kg year)1
per bear) was excreted in urine, 3% (1.1 ± 0.1 kg year)1
per bear) was egested in faeces and < 1% (0.3 ±
0.1 kg year)1 per bear) was retained in the body. On an
area basis, salmon-N redistribution rates were as high as
5.1 ± 0.7 mg m)2 per year per bear within 500 m of the
stream but declined with increasing distance. This level of
nutrient input may seem small but 15.5–17.8% of the total
N in spruce foliage within 500 m of the stream was
derived from salmon in this highly nutrient-limited system
(Hilderbrand et al. 1999). Moreover, the landscape-scale
effect of such input may be profound. For example, the
Tongass National Forest in Alaska contains nearly 5000
salmon-supporting streams. Forty-seven per cent of the
forested area within the Tongass falls within 0.5 km of a
salmon stream and over 90% within 5 km. The influx of
salmon-based nutrients through predators thus effectively
extends the interface between ocean and land, thereby
expanding the surface area over which ecological exchanges
take place (Ben-David et al. 1998). Salmon feeding by bears
also elevates total inorganic N pools (NH4+ + NO3))
threefold, and gaseous N2O flux by 32-fold compared
to riparian areas with lower bear activity (Holtgrieve et al.
2009). This level of nitrogen input provides on average
c. 2.2 g N m)2 to the riparian zone and constitutes up to
24% of the total riparian N budget (Helfield & Naiman
2006), a level that approaches fertilization inputs needed to
sustain silvicultural activities in the same kind of forest
ecosystem (Quinn et al. 2009).
2010 Blackwell Publishing Ltd/CNRS
1204 O. J. Schmitz, D. Hawlena and G. C. Trussell
Predator species living in large groups disperse nutrients
more evenly across landscapes. Seabirds that prey on
marine fish excrete ingested nutrients into terrestrial
breeding colonies (for examples, see Mizota 2009), creating
a nutrient source that varies with colony size. Moreover,
sustained guano input can saturate soils within the colony,
which leads to nutrient diffusion beyond the immediate
boundary of the colony, thereby creating a nutrient supply
gradient across the terrestrial landscape (Barrett et al. 2005;
Croll et al. 2005; Fukami et al. 2006; Maron et al. 2006).
Similarly, loggerhead sea turtles (Caretta caretta) deposit eggs
built from marine prey, such as jellyfish, into nests across
sandy beaches (Bouchard & Bjorndal 2000; Hannan et al.
2007). This activity contributes an average 3 g N m)2 and
0.3 g P m)2 to these nutrient-poor habitats after subtracting the nutrients that are returned to the sea by loggerhead
turtle hatchlings (Bouchard & Bjorndal 2000). The
temporally pulsed and spatially homogenous flux of
nutrients from marine ecosystems into freshwater ecosystems by salmon (as predators) mass migrations is well
chronicled (Gende et al. 2002; Naiman et al. 2002). A large
run of 20 million sockeye to the Bristol Bay region of
Alaska (Gende et al. 2002) can deliver as much as
5.4 · 107 kg of body tissue for the slow-cycle pathway
upon death after spawning, yielding 2.4 · 105 kg of P,
18 · 106 kg of N and 2.7 · 106 kg of Ca, plus other
macroelements to riparian zones. This nutrient delivery is
equivalent to the amount of fertilizer used to support
56 000 ha of intensive corn production in the US Midwest.
Alewives represent another example of en masse migration
from marine ecosystems to freshwater spawning sites that
lead to a temporally pulsed, and spatially homogenous,
influx of nutrients to freshwater ecosystems (Post &
Walters 2009). Finally, many predators, such as marine fish,
migrate daily from foraging areas (the pelagic zone) to
communal roosting areas (coral reefs). These migrations
can result in a 30–48% increase in NH4+, a 41–59% rise in
particulate N and a 68–94% rise in particulate P loadings
from the pelagic realm onto the reef (Meyer & Schultz
1985). Finally, crows that feed in urban areas and roost in
adjacent forests transport c. 2.28 kg ha)1 year)1 of P and
22.76 kg ha)1 year)1 of N to these forests (Fujita & Koike
2009). Thus, in urban forests with crow roosts, birds
contribute 2.6 times more allochthonous P and 0.66 times
more N than other pathways; whereas in urban forests
without roosts they contribute only 0.04 times more
allochthonous P and 0.013 times more N than other
pathways (Fujita & Koike 2009). These examples show
that predators that mediate nutrient translocation can
homogenize nutrients within the ecosystem to which they
transport nutrients. However, they may also increase
among-ecosystem nutrient heterogeneity, if some ecosystems receive the flux and others do not.
2010 Blackwell Publishing Ltd/CNRS
Idea and Perspective
Newly identified mechanisms
The classic mechanisms involve consumptive effects of
predators on prey, and other than trophic cascades, the
effects of predators on the release or translocation of
nutrients are largely direct. The newly identified mechanisms
also involve direct consumptive effects (mechanisms 5 and 6;
Table 1). However, others (mechanisms 7 and 8; Table 1)
involve exclusively non-consumptive predator effects that
lead to important indirect control of nutrient dynamics.
Decoupling carcass distribution from live-prey distribution
In some habitats, prey species tend to be disproportionately
vulnerable to predation leading to differences in the overall
distribution of live prey and locations where they tend to be
killed across the landscape (Kauffman et al. 2007). These kill
sites in turn may become nutrient ÔhotspotsÕ (Bump et al.
2009). Once wolves (Canis lupus) dispatch their moose (Alces
alces) prey, the kill site receives a high flux of N, P and K
into the slow-cycle pathway that becomes evident as a
100–600% increase in soil nutrients. This input translates
into a 25–47% (14–28 months postmortem) increase in
mean foliar nitrogen and a 25% increase in mean foliar
quality (measured as decline in C : N ratio) during the first
three growing seasons postmortem (Bump et al. 2009).
Wolf-killed moose were 12 times more common than
starvation-killed moose and the distribution of wolf-killed
moose showed a striking degree of clustering at the island
scale; the likelihood that such clustering resulted from
random chance was 0.1%. This example also illustrates the
different carcass distribution patterns that might emerge
between predator effects that are largely consumptive
(clumping in risky habitats) vs. non-consumptive (potential
clumping in refuges).
Transportation of prey carcasses away from hunting sites
to feeding sites (i.e. central place foraging) has the potential
to substantially redistribute nutrients across landscapes. For
example, bears can distribute marine-derived nutrients to
the surrounding forest by carrying 42–68% of the salmon
they kill away from streams (Quinn et al. 2009). Such
behaviour is also evident in the solitary hunting wading bird
(the limpkin, Aramus guarauna), which preys on wetland
snails that are evenly distributed across open water within
the wetland (Macek et al. 2009). Up to 80% of captured
snails are carried to and consumed within local patches of
emergent vegetation. Nutrients derived from empty snail
shells and unconsumed snail tissue can enhance levels of
total plant biomass and aboveground N and P in plants by
5· relative to control plots without limpkins. Stable isotope
analyses confirmed that snails are the dominant source of
nutrients to these patches, which cover 16% of the area and
create marked heterogeneity in productivity and patch
structure across the wetland landscape (Macek et al. 2009).
Idea and Perspective
Alteration of prey nutrient transport and release via prey capture
Predators are most notable for their functional role in
controlling prey abundance. They can have large effects on
ecosystems via systematic elimination of prey species. Such
strong effects are especially evident on islands having prey
that are important cross-ecosystem nutrient vectors but are
driven to local extinction by invasive predators. The
extinction of seabird breeding colonies by invading arctic
foxes (Alopex lagopus) and rats (Rattus spp.) (e.g. Croll et al.
2005; Fukami et al. 2006; Maron et al. 2006; Wardle et al.
2007, 2009) completely eliminates the influx of new guano
nutrients. In some cases (e.g. Croll et al. 2005; Maron et al.
2006), the reduction in guano input from 361.9 to 5.7 g m)2
(Croll et al. 2005) dramatically changes the composition of
the plant community which, in turn, transforms the entire
ecosystem into a new type. In an interesting twist of fate, the
elimination of seabirds by rats has altered the physical
structure of the soil (elimination of nesting burrows) to such
a great extent that the ability of soil microorganisms and
plants to sequester C is actually enhanced (Wardle et al. 2007).
Predators may also control the number of nutrient
vectors leaving or entering an ecosystem (Hodkinson et al.
2001; Ngai & Srivastava 2006). The presence of predatory
damsefly larvae (Mecistogaster modesta) living in wells of tankforming bromeliads leads to a 9.5-fold increase in the
retention of N from litter decomposition within the wells
that enhances the fertilization of bromeliad leaves. In the
absence of damselfly larvae, detritivorous insect prey are
able to leave the bromeliad tanks and thus export nutrients
by carrying litter-derived N with them. When damsefly
larvae are present, their consumption of detritivorous
insects reduces the export of N and instead converts this
mobile pool of N into faecal pellets that can be readily
decomposed by microbes or leached in a form of N that
becomes available to the bromeliad (Ngai & Srivastava
2006). Similarly, web-building spiders are the earliest
colonizers of newly exposed moraine substrates on glacier
forelands. These spiders entrap passing chironomid midges
that would otherwise pass over these sites, thereby creating
an allochtonous nutrient input to the moraine ecosystem
(Hodkinson et al. 2001).
Spatial redistribution of nutrients via predator-induced
prey habitat shift
Evolutionarily, it makes little sense for prey to passively
submit to capture by their predators so they often seek
refuge habits that reduce or eliminate predation risk. By
causing prey to undergo habitat shifts, predators can thus
have non-consumptive indirect effects on ecosystem processes. Early evidence for predator non-consumptive effects
on communities came from research in freshwater ecosystems (Kitchell et al. 1979; Carpenter & Kitchell 1996). This
and later work (Schindler et al. 1993) showed that diel
Predator control of ecosystem nutrient dynamics 1205
vertical migration of Daphnia in response to diel variation in
predation risk may alter the availability of nutrients in the
water column.
Chemical cues (kairimones) of predatory fish can cause
benthic chironomids (Chironomus riparius) to spend less time
foraging at the sediment surface and more time hiding in
their burrows within benthic sediments (Stief & Holker
2006). By retreating into burrows in response to predator
cues, chironomids increased the amount of organic matter
that enters the sediment relative to fishless controls because
they consume food particles and defecate within their
burrows. This non-consumptive effect causes a fivefold
increase in the amount of organic matter within the
sediment layer to be decomposed and mineralized and later
taken up by aquatic vascular plants (Stief & Holker 2006).
Under conditions where chironomids do not face risk and in
treatment conditions without chironomids, organic matter
remained at the sediment surface and was broken down by
microbes that in turn released dissolved organic and
inorganic carbon and ammonium to the water column to
be taken up by microbes and algae (Stief & Holker 2006).
The conditions in the treatment without chironomids
resembles what might also be expected if predator effects
on chironomids were purely consumptive, i.e. a decline in
chironomid abundance should lead to the accumulation of
organic mater at the sediment surface. This suggests that
consumptive and non-consumptive predator effects should
lead to qualitatively different spatial distributions of organic
matter within the aquatic system. Predators that have
consumptive effects should have little net effect on the
organic matter decomposition and redistribution in this kind
of system, whereas predators causing non-consumptive
effects should cause redistribution of organic matter.
Prior to wolf introductions into Yellowstone National
Park, USA, ungulates foraged most intensively on grassland
sites with high primary productivity. Wolf introductions
altered the use of grazing land by ungulates (especially elk
Cervus elaphus and pronghorn Antilocapra americana) leading to
a 60–90% reduction in grazing impact at these sites (Frank
2008). This habitat shift is translated into a 50–60% decline
in forage N, P and macronutrient content, and N
mineralization rate.
Alteration of community composition and nutrient dynamics via
predator-induced herbivore foraging shifts
A generalist grasshopper herbivore (Melanoplus femurrubrum)
selects nutrients from a mixture of meadow grass (specifically Poa pratensis) and herb species (specifically the
competitively dominant goldenrod Solidago rugosa) (Schmitz
2006). Upon facing predation risk by a hunting spider
(Pisuarina mira), it seeks refuge and increases foraging in
structurally complex Solidago (Schmitz 2006). Experimental
manipulation of predator presence showed that the evasive
2010 Blackwell Publishing Ltd/CNRS
1206 O. J. Schmitz, D. Hawlena and G. C. Trussell
behaviour of grasshoppers facing risk altered Solidago
dominance over other plants, allowing other herb species,
which may otherwise have been intolerant of shady
conditions and low N supply caused by Solidago, to
proliferate (Schmitz 2006). This risk-mediated alteration in
the plant community led to a doubling of N-mineralization
rate (Schmitz 2006). Non-consumptive effects via foraging
activity shifts are also evident in marine sea grass system.
Here, predatory blue crabs (Callinectes sapidus) can directly
reduce amphipod and isopod grazer feeding activities
(a direct consumptive reduction in grazer density may also
be involved), thereby contributing to an increase in labile
organic matter at the sediment surface and a positive
response in sediment bacterial biomass (Canuel et al. 2007;
Spivak et al. 2007). This shift in bacterial community
composition can alter organic matter decomposition and
nutrient availability to plants (Spivak et al. 2009).
Where do we go from here?
Our coverage of the different identified mechanisms is
uneven because, in some instances (e.g. nutrient translocation associated with salmon runs), the highly abundant
nature of the species involved has allowed direct observation of effects. In other instances (e.g. any of the nonconsumptive mechanisms), resolution of the mechanisms
involved required explicit experimental manipulation of
predator abundance. This unevenness begs for further
research that explores whether certain mechanisms are
indeed idiosyncratic to particular ecosystem types or
predator species or more broadly representative. Furthermore, while existing modelling efforts, which were important motivators of research on links between trophic
interactions and ecosystem nutrient dynamics (e.g. DeAnglis
1992; Polis et al. 1997; Leroux & Loreau 2008), do recognize
that predators can drive nutrient dynamics, they only
effectively consider two of the eight mechanisms reported
here (i.e. trophic cascades and cross-ecosystem fluxes). Our
theoretical perspective on this topic thus needs to be
expanded considerably.
Although much of the evidence so far demonstrates a
predator effect, individual studies often do not quantify the
importance of the effect in terms of its contribution to the
entire nutrient budget. We have tried to estimate the
proportional contributions whenever the data allowed
calculations, but the overall significance of predators to
ecosystem productivity and trophic structure needs much
more systematic analysis. One attempt in this direction
involved estimating the nitrogen and phosphorus loading by
predatory waterbirds in Netherland wetlands using a
physiological modelling approach. Average external (i.e.
importing) loading estimates ranged from 38.1 to 91.5
tonnes N and 16.7 to 18.2 tonnes P per year, whilst internal
2010 Blackwell Publishing Ltd/CNRS
Idea and Perspective
(i.e. recycling) loading estimates ranged from 53.1 to 140.5
tonnes N and 25.2 to 39.2 tonnes P per year (Hahn et al.
2007). On a landscape scale, such nutrient loading by
carnivorous waterbirds is of minor importance for freshwater habitats in the Netherlands with 0.26–0.65 kg
N ha)1 year)1 and 0.12–0.16 kg P ha)1 year)1. However,
on a local scale, breeding colonies may be responsible for
significant P loading, leading again to spatial heterogeneity in
nutrient distribution and potential ecosystem productivity.
A similar attempt was made to estimate the influence of
terrestrial-borne P subsidy that enters a lake. In this case,
a bioenergetics model revealed that terrestrial-derived
P release from fish (Alburnus alburnus) that fed on terrestrial
insects may be important only in small oligotrophic lakes.
Increasing the perimeter-to-area ratio of the interface
between donor (terrestrial) and recipient (lake) habitats
causes the importance of subsidy to diminish and even
become negligible in large lakes (Mehner et al. 2005).
These examples highlight how the largely community
ecological perspective on trophic control could benefit from
more deliberate inclusion of an ecosystem-based methodology that traces and quantifies nutrient fluxes and total
ecosystem nutrient budgets (e.g. Pace et al. 2004). Uncovering indirect effects on ecosystem properties and functions
also requires more effective and creative integration of the
experimental tradition in community ecology, which manipulates or systematically compares functioning in areas where
predators are part of the landscape with those where they
are not (e.g. Frank 2008), with ecosystem approaches and
perspectives. Finally, it argues for expanding classic biogeochemical approaches in ecosystems (e.g. Schlesinger 1991)
to explicitly include organismal biology and ecology in analyses of nutrient cycling (e.g. see Pomeroy 2001; Sterner &
Elser 2002; Polis et al. 2004).
In particular, predation risk induces stress (Creel &
Christianson 2008) that can elevate respiration and shift
nutrient demand from N-rich proteins that support growth
and reproduction towards carbohydrates (C-rich, N-poor)
that fuel the heightened respiratory demands of antipredator
behaviour (Stoks et al. 2005; Trussell et al. 2006, 2008). This
means that C : N : P ratios within prey and in excreta or
egesta may also become altered by predator physiological
effects. Such plasticity in C : N : P uptake and body
composition in response to altered metabolic rate is
consistent with theoretical expectations of threshold elemental ratios (Frost et al. 2006) and suggests that our
understanding of physiological plasticity of consumers to
environmental changes must be improved if we are to
develop more robust predictions of nutrient dynamics in
natural systems (see also Hillebrand et al. 2008).
We focused here on examples that provide comparatively
strong evidence of predator effects on nutrient dynamics.
However, predator effects on ecosystem functions that are
Idea and Perspective
either weak or absent may also arise in both plant- and
detritus-based chains (Shurin et al. 2006; Schmitz 2010).
Because predators and herbivores also vary in traits such as
hunting mode (e.g. sit-&-wait, active pursuit) and feeding
mode (e.g. generalist or specialist leaf chewers, sap feeders,
leaf miners, etc.), the nature and strength of predator effects
may vary depending upon the particular combinations of
predator and herbivore feeding modes represented in the
community (Schmitz 2010). Predator effects may be
stronger in simple food webs and weaker when the prey
species pool is diverse; and may depend upon the degree of
coupling between plant- and detritus-based chains (Pomeroy
2001; Schmitz 2010). The magnitudes of ecosystem
responses may also depend upon whether predator controls
are direct or indirect. For example, predator nutrient
excretion may directly and strongly increase the availability
of inorganic nutrients to plants (fast cycling), whereas
indirect predator controls on plants may have a weaker
effect (slow cycling) on nutrient availability (Shurin et al.
2006) that requires more time to develop. Unfortunately, the
fact that there are only a few studies that have explored
predator effects on ecosystem functioning prevents definitive statements about which of these factors are most
important in modulating the strength of predator effects on
nutrient dynamics (Schmitz 2010). This deficiency is all the
more remarkable considering that Kitchell et al. (1979)
called for analyses of predator effects on ecosystem nutrient
dynamics 30+ years ago.
CONCLUSIONS
Much of our current conceptualization and empirical understanding of predator control of nutrient dynamics is based on
studies in freshwater ecosystems (Vanni 2002). Research in
aquatic systems has shown that predators can rapidly alter the
rate of nutrient cycling and ecosystem productivity via direct
excretion and nutrient translocation because they live within a
medium that can quickly dissolve and dissipate nutrients to
enhance the production of species (phytoplankton and
bacteria) with rapid life cycles (Carpenter et al. 1992; Vanni
2002). It has been hypothesized that these properties may
explain why aquatic (especially pelagic) ecosystems are more
strongly regulated by top-down control compared with
benthic and terrestrial systems (Shurin et al. 2006). However,
conception of this hypothesis may largely reflect the fact that
empirical studies on the link between predators and nutrient
dynamics are predominantly from aquatic systems. The
accumulation of recent evidence from other ecosystem types
(Table 2) shows that predator effects on nutrient dynamics
may be ubiquitous and operate within and across ecosystem
boundaries (Table 2).
The insights provided here also call for evolution in
thinking about the conservation value of predator species.
Predator control of ecosystem nutrient dynamics 1207
Prevailing attitudes have been very species or populationcentric, with predators being valued for their charismatic
identity or for their ability to control prey that would
otherwise become pests. We suggest that predator effects
can have multifarious direct and indirect effects on
ecosystem nutrient cycling and, accordingly, play a pivotal
role in the provisioning of ecosystem services. This view
reinforces pleas for taking a more whole ecosystem perspective when devising conservation strategies (Sinclair & Byrom
2006; Schmitz 2010).
ACKNOWLEDGEMENTS
We thank three anonymous referees for helpful comments.
This work was supported by Grants from the National
Science Foundation to O.J. Schmitz (DEB-0816504) and
G.C. Trussell (OCE-0648525, OCE-0727628). This is
contribution No. 267 from the Marine Science Center.
REFERENCES
Barrett, K., Anderson, W.B., Wait, D.A., Grismer, L.L., Polis,
G.A. & Rose, M.D. (2005). Marine subsidies alter the diet and
abundance of insular and coastal lizard populations. Oikos, 109,
145–153.
Ben-David, M., Hanley, T.A. & Schell, D.M. (1998). Fertilization of
terrestrial vegetation by spawning Pacific salmon: the role of
flooding and predator activity. Oikos, 83, 47–55.
Bouchard, S.S. & Bjorndal, K.A. (2000). Sea turtles as biological
transporters of nutrients and energy from marine to terrestrial
ecosystems. Ecology, 81, 2305–2313.
Breuning-Madsen, H., Ehlers, C.B. & Borggard, O.K. (2008). The
impact of perennial cormorant colonies on soil phosphorus
status. Geoderma, 148, 51–54.
Bump, J.K., Webster, C.R., Vucetich, J.A., Peterson, R.O., Shields,
J.M. & Powers, M.D. (2009). Ungulate carcasses perforate ecological filters and create biogeochemical hotspots in forest herbaceous layers allowing trees a competitive advantage. Ecosystems,
12, 996–1007.
Canuel, E.A., Spivak, A.C., Waterson, E.J. & Duffy, J.E. (2007).
Biodiversity and food web structure influence short-term accumulation of sediment organic matter in an experimental seagrass
system. Limnol. Oceanogr., 52, 590–602.
Carpenter, S.R. & Kitchell, J.F. (1988). Consumer control of lake
productivity. Bioscience, 38, 764–769.
Carpenter, S.R. & Kitchell, J.F. (1996). The Trophic Cascade in Lakes.
Cambridge Universit. Press, Cambridge.
Carpenter, S.R., Cottingham, K.L. & Schindler, D.E. (1992). Biotic
feedbacks in lake phosphorus cycles. TREE, 7, 332–336.
Costanza, R., dArge, R., deGroot, R., Farber, S., Grasso, M. &
Hannon, B (1997). The value of the worldÕs ecosystem services
and natural capital. Nature, 387, 253–260.
Crait, J.R. & Ben-David, M. (2007). Effects of river otter activity on
terrestrial plants in trophically altered Yellowstone Lake. Ecology,
88, 1040–1052.
Creel, S. & Christianson, D. (2008). Relationships between direct
predation and risk effects. TREE, 23, 194–201.
2010 Blackwell Publishing Ltd/CNRS
1208 O. J. Schmitz, D. Hawlena and G. C. Trussell
Croll, D.A., Maron, J.L., Estes, J.A., Danner, E.M. & Byrd, G.V.
(2005). Introduced predators transform subarctic islands from
grassland to tundra. Science, 307, 1959–1961.
DeAnglis, D.L. (1992). Dynamics of Nutrient Cycling and Food Webs.
Chapman and Hall, London.
Dobson, A., Lodge, D., Alder, J., Cumming, G.S., Keymer, J. &
McGlade, J. et al. (2006). Habitat loss, trophic collapse, and the
decline of ecosystem services. Ecology, 87, 1915–1924.
Duffy, J.E. (2003). Biodiversity loss, trophic skew and ecosystem
functioning. Ecol. Lett., 6, 680–687.
Dunham, A.E. (2008). Above and below ground impacts of terrestrial mammals and birds in a tropical forest. Oikos, 117, 571–
579.
Frank, D.A. (2008). Evidence for top predator control of a grazing
ecosystem. Oikos, 117, 1718–1724.
Frost, P.C., Bemstead, J.P., Cross, W.F., Hillebrand, H., Larson,
J.H. & Xenopoulos, M.A. et al. (2006). Threshold elemental ratios of carbon and phosphorus in aquatic consumers. Ecol. Lett.,
9, 774–779.
Fujita, M. & Koike, F. (2009). Landscape effects on ecosystems:
birds as active vectors of nutrient transport to fragmented
urban forests versus forest-dominated landscapes. Ecosystems, 12,
391–400.
Fukami, T., Wardle, D.A., Bellingham, P.J., Mulder, C.P.H.,
Towns, D.R. & Yeates, G.W. et al. (2006). Above- and belowground impacts of introduced predators in seabird-dominated
island ecosystems. Ecol. Lett., 9, 1299–1307.
Gende, S.M., Edwards, R.T., Willson, M.F. & Wipfli, M.S. (2002).
Pacific salmon in aquatic and terrestrial ecosystems. Bioscience,
52, 917–928.
Gende, S.M., Miller, A.E. & Hood, E. (2007). The effects of salmon carcasses on soil nitrogen pools in a riparian forest of
southeastern Alaska. Can. J. For. Res. – Rev. Can. Rech. For., 37,
1194–1202.
Hahn, S., Bauer, S. & Klaassen, M. (2007). Estimating the contribution of carnivorous waterbirds to nutrient loading in freshwater habitats. Freshw. Biol., 52, 2421–2433.
Halpern, B.S., Borer, E.T., Seabloom, E.W. & Shurin, J.B. (2005).
Predator effects on herbivore and plant stability. Ecol. Lett., 8,
189–194.
Hannan, L.B., Roth, J.D., Ehrhart, L.M. & Weishampel, J.F.
(2007). Dune vegetation fertilization by nesting sea turtles.
Ecology, 88, 1053–1058.
Heithaus, M.R., Frid, A., Wirsing, A.J. & Worm, B. (2008). Predicting ecological consequences of marine top predator declines.
TREE, 23, 202–210.
Helfield, J.M. & Naiman, R.J. (2006). Keystone interactions:
salmon and bear in riparian forests of Alaska. Ecosystems, 9,
167–180.
Hilderbrand, G.V., Hanley, T.A., Robbins, C.T. & Schwartz, C.C.
(1999). Role of brown bears (Ursus arctos) in the flow of marine
nitrogen into a terrestrial ecosystem. Oecologia, 121, 546–550.
Hillebrand, H., Frost, P. & Liess, A. (2008). Ecological stoichiometry of indirect grazer effects on periphyton nutrient content.
Oecologia, 155, 619–630.
Hobara, S., Koba, K., Osono, T., Tokuchi, N., Ishida, A. &
Kameda, K. (2005). Nitrogen and phosphorus enrichment and
balance in forests colonized by cormorants: Implications of the
influence of soil adsorption. Plant and Soil, 268, 89–101.
2010 Blackwell Publishing Ltd/CNRS
Idea and Perspective
Hodkinson, I.D., Coulson, S.J., Harrison, J. & Webb, N.R. (2001).
What a wonderful web they weave: spiders, nutrient capture and
early ecosystem development in the high Arctic – some counterintuitive ideas on community assembly. Oikos, 95, 349–352.
Holtgrieve, G.W., Schindler, D.E. & Jewett, P.K. (2009). Large
predators and biogeochemical hotspots: brown bear (Ursus arctos) predation on salmon alters nitrogen cycling in riparian soils.
Ecol. Res., 24, 1125–1135.
Kauffman, M.J., Varley, N., Smith, D.W., Stahler, D.R.,
MacNulty, D.R. & Boyce, M.S. (2007). Landscape heterogeneity
shapes predation in a newly restored predator-prey system.
Ecol. Lett., 10, 690–700.
Kitchell, J.F., OÕNeill, R.V., Webb, D., Gallepp, G.W., Bartell,
S.M. & Koonce, J.F. et al. (1979). Consumer regulation of
nutrient cycling. Bioscience, 29, 28–34.
Leroux, S.J. & Loreau, M. (2008). Subsidy hypothesis and strength
of trophic cascades across ecosystems. Ecol. Lett., 11, 1147–
1156.
Macek, P., Rejmankova, E. & Fuchs, R. (2009). Biological activities
as patchiness driving forces in wetlands of northern Belize.
Oikos, 118, 1687–1694.
Maron, J.L., Estes, J.A., Croll, D.A., Danner, E.M., Elmendorf,
S.C. & Buckelew, S.L. (2006). An introduced predator alters
Aleutian Island plant communities by thwarting nutrient subsidies. Ecol. Monogr., 76, 3–24.
McNaughton, S.J., Ruess, R.W. & Seagle, S.W. (1988). Large
mammals and process dynamics in African ecosystems. Bioscience,
38, 794–800.
Mehner, T., Ihlau, J., Dorner, H. & Holker, F. (2005). Can feeding
of fish on terrestrial insects subsidize the nutrient pool of lakes?.
Limnol. Oceanogr., 50, 2022–2031.
Meyer, J.L. & Schultz, E.T. (1985). Migrating haemulid fishes as
a source of nutrients and organic matter on reefs. Limnol.
Oceanogr., 30, 146–156.
Mikola, J. & Setala, H. (1998). No evidence of trophic cascades
in an experimental microbial-based soil food web. Ecology, 79,
153–164.
Mizota, C. (2009). Nitrogen isotopic patterns of vegetation as
affected by breeding activity of Black-tailed Gull (Larus
crassiostris): a coupled analysis of feces, inorganic soil nitrogen
and flora. Appl. Geogr., 24, 2027–2033.
Moore, J.C., Berlow, E.L., Coleman, D.C., DeReuter, P.C.,
Dong, Q. & Hostings, A. et al. (2004). Detritus, trophic dynamics
and biodiversity. Ecol. Lett., 7, 584–600.
Myers, R.A. & Worm, B. (2003). Rapid worldwide depletion of
predatory fish communities. Nature, 423, 280–283.
Naiman, R.J., Bilby, R.E., Schindler, D.E. & Helfield, J.M. (2002).
Pacific salmon, nutrients, and the dynamics of freshwater and
riparian ecosystems. Ecosystems, 5, 399–417.
Ngai, J.T. & Srivastava, D.S. (2006). Predators accelerate nutrient
cycling in a bromeliad ecosystem. Science, 314, 963.
Osono, T., Hobara, S., Koba, K., Kameda, K. & Takeda, H. (2006).
Immobilization of avian excreta-derived nutrients and reduced
lignin decomposition in needle and twig litter in temperate
coniferous forest. Soil Biology & Biochemistry, 38, 517–525.
Pace, M.L., Cole, J.J., Carpenter, S.R., Kitchell, J.F., Hodgson,
J.R. & Van de Bogert, M.C. (2004). Whole-lake carbon-13
additions reveal terrestrial support of aquatic food webs. Nature,
427, 240–243.
Idea and Perspective
Pauly, D., Christensen, V., Dalsgaard, J., Froese, R. & Torres, F.
(1998). Fishing down marine food webs. Science, 279, 860–863.
Persson, A. & Svensson, J.M. (2006). Effects of benthivorous fish
on biogeochemical processes in lake sediments. Freshw. Biol., 51,
1298–1309.
Polis, G.A., Anderson, W.B. & Holt, R.D. (1997). Toward an integration of landscape and food web ecology: the dynamics of
spatially subsidized food webs. Annu. Rev. Ecol. Syst., 28, 289–316.
Polis, G.A., Power, M.E. & Huxel, G.R. (2004). Food Webs at the
Landscape Level. University of Chicago Press, Chicago.
Pomeroy, L.R. (2001). Caught in the food web: complexity made
simple? Sci. Mar., 65, 31–40.
Post, D.M. & Walters, A.W. (2009). Nutrient excretion rates of
anadromous Alewives during their spawning migration. T. Am.
Fish. Soc., 138, 264–268.
Quinn, T.P., Carlson, S.M., Gende, S.M. & Rich, H.B. (2009).
Transportation of Pacific salmon carcasses from streams to
riparian forests by bears. Can. J. Zool., 87, 195–203.
Ritchie, M.E., Tilman, D. & Knops, J.M.H. (1998). Herbivore
effects on plant and nitrogen dynamics in oak savanna. Ecology,
79, 165–177.
Schindler, D.E., Kitchell, J.F., He, X., Carpenter, S.R., Hodgson,
J.R. & Cottingham, K.L. (1993). Food web structure and
phosphorous cycling in lakes. T. Am. Fish. Soc., 122, 756–772.
Schlesinger, W.H. (1991). Biogeochemistry: An Analysis of Global
Change. Academic Press, San Diego.
Schmitz, O.J. (2006). Predators have large effects on ecosystem
properties by changing plant diversity, not plant biomass.
Ecology, 87, 1432–1437.
Schmitz, O.J. (2008). Herbivory from individuals to ecosystems.
Annu. Rev. Ecol. Syst., 39, 133–152.
Schmitz, O.J. (2010). Resolving Ecosystem Complexity. Princeton
University Press, Princeton.
Sergio, F., Newton, I., Marchesi, L. & Pedrini, P. (2006). Ecologically justified charisma: preservation of top predators delivers
biodiversity conservation. J. Appl. Ecol., 43, 1049–1055.
Shurin, J.B., Gruner, D.S. & Hillebrand, H. (2006). All wet or dried
up?. Real differences between aquatic and terrestrial food webs
Proc. R. Soc. Lond., B, Biol. Sci., 273, 1–9.
Sin, H., Beard, K.H. & Pitt, W.C. (2008). An invasive frog, Eleutherodactylus coqui, increases new leaf production and leaf litter
decomposition rates through nutrient cycling in Hawaii. Biol.
Invasions, 10, 335–345.
Sinclair, A.R.E. & Byrom, A.E. (2006). Understanding ecosystem
dynamics for conservation of biota. J. Anim. Ecol., 75, 64–79.
Soule, M.E., Estes, J.A., Miller, B. & Honnold, D.L. (2005).
Strongly interacting species: conservation policy, management,
and ethics. Bioscience, 55, 168–176.
Predator control of ecosystem nutrient dynamics 1209
Spivak, A.C., Canuel, E.A., Duffy, J.E. & Richardson, J.P. (2007).
Top-down and bottom-up controls on sediment organic matter
composition in an experimental seagrass ecosystem. Limnol.
Oceanogr., 52, 2595–2607.
Spivak, A.C., Canuel, E.A., Duffy, J.E., Douglass, J.G. &
Richardson, J.P. (2009). Epifaunal community composition and
nutrient addition alter sediment organic matter composition in
a natural eelgrass Zostera marina bed: a field experiment.
Mar. Ecol. Prog. Ser., 376, 55–67.
Sterner, R.W. & Elser, J.J. (2002). Ecological Stoichiometry. Princeton
University Press, Princeton.
Stief, P. & Holker, F. (2006). Trait-mediated indirect effects of
predatory fish on microbial mineralization in aquatic sediments.
Ecology, 87, 3152–3159.
Stoks, R., De Block, M. & McPeek, M.A. (2005). Alternative
growth and energy storage responses to mortality threats in
damselflies. Ecol. Lett., 8, 1307–1316.
Thebault, E. & Loreau, M. (2006). The relationship between biodiversity and ecosystem functioning in food webs. Ecol. Res., 21,
17–25.
Trussell, G.C., Ewanchuk, P.J. & Matassa, C.M. (2006). The fear of
being eaten reduces energy transfer in a simple food chain.
Ecology, 87, 2979–2984.
Trussell, G.C., Ewanchuk, P.J. & Matassa, C.M. (2008). Resource
identity modifies the influence of predation risk on ecosystem
function. Ecology, 89, 2798–2807.
Vanni, M.J. (2002). Nutrient cycling by animals in freshwater
ecosystems. Annu. Rev. Ecol. Syst., 33, 341–370.
Wardle, D.A. & Bardgett, R.D. (2004). Human-induced changes in
large herbivorous mammal density: the consequences for
decomposers. Front. Ecol. Environ., 2, 145–153.
Wardle, D.A., Bellingham, P.J., Fukami, T. & Mulder, C.P.H.
(2007). Promotion of ecosystem carbon sequestration by invasive predators. Biol. Lett., 3, 479–482.
Wardle, D.A., Bellingham, P.J., Bonner, K.I. & Mulder, C.P.H.
(2009). Indirect effects of invasive predators on litter decomposition and nutrient resorption on seabird-dominated islands.
Ecology, 90, 452–464.
Wootton, J.T. & Emmerson, M. (2005). Measurement of interaction strength innature. Annu. Rev. Ecol. Syst., 36, 419–444.
Editor, Helmut Hillebrand
Manuscript received 11 March 2010
First decision made 5 April 2010
Second decision made 1 June 2010
Manuscript accepted 3 June 2010
2010 Blackwell Publishing Ltd/CNRS