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

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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

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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

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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

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(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.

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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.

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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

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