Food web structure in riverine landscapes
←
→
Page content transcription
If your browser does not render page correctly, please read the page content below
Freshwater Biology (2002) 47, 777–798
Food web structure in riverine landscapes
G U Y W O O D W A R D * and A L A N G . H I L D R E W †
*Institute of Ecology and Resource Management, University of Edinburgh, Edinburgh, U.K.
†School of Biological Sciences, Queen Mary University of London, London, U.K.
SUMMARY
1. Most research on freshwater (and other) food webs has focused on apparently discrete
communities, in well-defined habitats at small spatial and temporal scales, whereas in
reality food webs are embedded in complex landscapes, such as river corridors. Food web
linkages across such landscapes may be crucial for ecological pattern and process,
however. Here, we consider the importance of large scale influences upon lotic food webs
across the three spatial dimensions and through time.
2. We assess the roles of biotic factors (e.g. predation, competition) and physical habitat
features (e.g. geology, land-use, habitat fragmentation) in moulding food web structure at
the landscape scale. As examples, external subsidies to lotic communities of nutrients,
detritus and prey vary along the river corridor, and food web links are made and broken
across the land–water interface with the rise and fall of the flood.
3. We identify several avenues of potentially fruitful research, particularly the need to
quantify energy flow and population dynamics. Stoichiometric analysis of changes in
C : N : P nutrient ratios over large spatial gradients (e.g. from river source to mouth, in
forested versus agricultural catchments), offers a novel method of uniting energy flow and
population dynamics to provide a more holistic view of riverine food webs from a
landscape perspective. Macroecological approaches can be used to examine large-scale
patterns in riverine food webs (e.g. trophic rank and species–area relationships). New
multivariate statistical techniques can be used to examine community responses to
environmental gradients and to assign traits to individual species (e.g. body-size,
functional feeding group), to unravel the organisation and trophic structure of riverine
food webs.
Keywords: energy subsidies, macroecology, spatial scale, stable isotopes, stoichiometry
scales, usually due to logistic constraints. Such approa-
Introduction
ches may be inappropriate, however, because species
The notion that linking consumers and resources in a in the web occupy, or interact over, a much greater
community via a network of trophic interactions, the area than is represented in the study. A similar
food web, can reveal fundamental properties of a mismatch may occur between the temporal dynamics
system is one of the earliest ideas in ecology (Elton, of communities and the time-span of research. Land-
1927; MacArthur, 1955). However, despite the early scape ecology, on the other hand, has been carried out
development of this field, good quality food web data mostly at much larger scales, often at the catchment
are still scarce (Cohen et al., 1993; Hall & Raffaelli, scale or above [although the ‘landscape’ approach can
1993). Most food web studies are carried out at small be applied at any scale (Wiens, 2002)], but rarely
within the context of food web theory. Consequently,
Correspondence: Dr Guy Woodward, Institute of Ecology and we have drawn upon literature from two disciplines
Resource Management, University of Edinburgh, Darwin which, although they have obvious common ground,
Building, Mayfield Road, Edinburgh EH9 3JU, U.K. are rarely integrated by their respective practitioners
E-mail: guy.woodward@ed.ac.uk (but see Polis, Anderson & Holt, 1997). Here we aim to
Ó 2002 Blackwell Science Ltd. 777778 G. Woodward and A.G. Hildrew
redress this gap by considering riverine food webs lotic (and other) food webs has been carried out at
from a landscape perspective, although the paucity of small scales (Levin, 1992). There is a strong spatial
suitable large scale studies has necessitated a good bias towards small streams (e.g. Tavares-Cromar &
deal of extrapolation. Williams, 1996; Schmid-Araya & Schmid, 2000) and
Several important recent advances in food web, even towards small scales within these systems, as in
landscape and freshwater ecology suggest that a most predator/prey experiments (e.g. Englund, 1997),
consideration of riverine food webs at larger scales although a few studies of trophic interactions have
(i.e. at the reach scale and above) is overdue. These been performed recently at the reach scale or above
points, summarised below, are discussed in greater (e.g. Lancaster, 1996; Crowl et al., 1997; Wallace et al.,
detail later, where we attempt to create a more unified 1999). Similarly, large-scale energy fluxes linking food
view of riverine webs and to identify fruitful direc- webs in different ecosystems have been largely
tions for future research. We address the need to ignored, until relatively recently (Polis et al., 1997;
produce more quantitative, rather than qualitative, Flory & Milner, 1999; Milner et al., 2000; Wipfli &
food webs (Cohen et al., 1993; Hall & Raffaelli, 1993), Gregovich, in press). There is also a marked temporal
and to do so in a way that combines population bias towards the study of truncated ‘snapshots’, both
dynamics and energy flow. Potentially, food web empirically [via gut contents analysis (e.g. Allan,
stoichiometry provides such an integrated approach 1982)] and experimentally [via short-term field
to understanding riverine webs, particularly at large manipulations (e.g. Cooper, Walde & Peckarsky,
spatial scales (e.g. changes in C : N : P ratios between 1990)]. Few studies have addressed seasonal variation
a river’s source and its mouth). Community assembly in riverine webs (e.g. Closs & Lake, 1994; Tavares-
models have stressed the importance of ‘external’ Cromar & Williams, 1996; Thompson & Townsend,
physical factors in determining the regional species 1999; Schmid-Araya et al., in press), let alone changes
pool, within which the ‘internal’ dynamics of the food across years (but see Lancaster et al., 1996; Wood-
web operate to create the realised pool from those ward, Jones & Hildrew, in press). Although a few
species available (Belyea & Lancaster, 1999). We recent studies have examined the intergenerational
examine the respective roles of physical and biotic population dynamics of predators and prey within
factors, and how these interact among the three lotic webs (Peckarsky, Cooper & McIntosh, 1997;
spatial dimensions and across time to mould the Speirs et al., 2000), empirical data are still scarce. Most
riverine food webs that we find in nature. Other mathematical models of food webs deal with inter-
potentially important landscape scale effects include generational dynamics (e.g. May, 1972; Pimm, 1980;
the roles of dispersal, colonisation and cross-system McCann, Hastings & Huxel, 1998) and it is often at
energy subsidies. By applying a landscape perspective this temporal scale that indirect effects are manifested,
we may gain clearer insight into how energy flow and yet there are virtually no long-term data with which to
population dynamics shape web pattern and process. validate these models (Hall & Raffaelli, 1993; May,
Our ultimate objective was to draw these currently 1999).
disparate themes and novel approaches together to
promote a more coherent and spatially explicit syn-
Integrating from above and below the landscape scale
thesis of food webs in the riverine landscape.
Riverine food webs are produced by forces that
impinge upon them at a range of spatiotemporal
Scale and food web structure scales (Fig. 1), although patterns that emerge at one
scale can be produced by processes that operate at
Scale dependence: putting empiricism, experiments
another (Levin, 1992; Peckarsky et al., 1997). It has
and models into the same space
been proposed that lotic communities are structured
Despite the fact that ecologists have increasingly as nested hierarchies, with each stage forcing or
stressed the importance of scale (e.g. Levin, 1992; constraining the stage below it (Frissell et al., 1986;
Cooper et al., 1998), remarkably little attention has Hildrew & Giller, 1992; Ward, 1998). Thus, processes
been paid to processes that influence food webs at operating at the landscape scale (e.g. dispersal of
large spatial and temporal scales. Most research on adult insects across catchments) can shape food web
Ó 2002 Blackwell Science Ltd, Freshwater Biology, 47, 777–798Riverine food webs 779 Fig. 1 The spatiotemporal scaling of riverine food webs. Selected fields of investigation that might be expected to affect web structure and processes at the different scales are highlighted as text. Solid double-headed arrows indicate the typical spatiotemporal limits of these investigations; the dashed arrows indicate rarer instances, where these limits are exceeded. structure at smaller scales (e.g. the species pool within scales are often autocorrelated (Wiens, 1989; Ward, a stream). However, processes at smaller scales can 1997), when shifting our perspective in one dimension also produce patterns at larger scales. For instance, we need to adjust the other dimension accordingly although aggregative responses of predators to prey (Fig. 1). produce positive associations at small scales (e.g. At the small scale typical of field experiments, Hildrew & Townsend, 1982), the inverse is often behavioural interactions, mobility and patchiness in apparent at larger scales, as the effects of predation resource availability become the important factors have been ‘filtered’ through time and space. At these that determine predator impacts and local food web larger scales the longer-term consequences of preda- structure (Lancaster, Hildrew & Townsend, 1991; Sih tion upon population size become evident [e.g. neg- & Wooster, 1994; Englund, 1997). For example, the ative correlations between fish and large invertebrate crowding of predators and prey into flow refugia predators across streams (Hildrew, Townsend & during spates can lead to pulses of strong predation in Francis, 1984)]. Clearly, because temporal and spatial these patches (e.g. Lancaster, 1996). Moving up the Ó 2002 Blackwell Science Ltd, Freshwater Biology, 47, 777–798
780 G. Woodward and A.G. Hildrew
spatiotemporal scale, however, the emphasis shifts to Hughes et al., 1998; Wilcock, Hildrew & Nichols,
longer-term population dynamics (Englund, 1997; 2001). These results are not necessarily contradictory,
Cooper et al., 1998). Obviously, the scale will also however. Rare dispersal events between streams may
vary with the taxonomy of the organisms examined: be sufficient to homogenise local populations gen-
populations of rotifers can grow and shrink through etically, if the effective population size of breeding
many generations in a short time and over areas females is very small. Such dispersal events might
considerably smaller than a single square metre, not be important in terms of population dynamics,
whereas fish or bird populations can cover many except where local extinction has occurred. Because
square kilometres (Winemiller & Jepsen, 1998). This of high fecundity, even a tiny number of gravid
tends to lead to a trophic level bias at different spatial female colonists from either the natant or a neigh-
scales, as small species (e.g. microcrustacea) tend to be bouring stream may be sufficient to repopulate an
lower in the web than large species, and body-size is entire system, particularly if larval survival is
an important determinant of web structure (Warren & strongly density-dependent, as appears to be true
Lawton, 1987; Woodward & Hildrew, 2001). Indeed, for many insects (Hopper, Crowley & Kielman, 1996;
top predators that require large territories may link Ray & Hastings, 1996). Such a population bottleneck
several different webs together (Hall & Raffaelli, in each generation could essentially ‘reset the clock’
1993). However, as most lotic food webs have been of the outcomes of predatory or competitive interac-
constructed with an emphasis upon macroinverte- tions among the larvae. Because only a few colonists
brates (e.g. Hildrew, Townsend & Hasham, 1985; would have to survive as adults, dispersal could
Closs & Lake, 1994; Tavares-Cromar & Williams, 1996; enable populations to persist over large temporal
Thompson & Townsend, 1999), for convenience sake and spatial scales, and link apparently isolated food
we consider population level effects to occur primar- webs across the riverine landscape. Indeed, many
ily between the scales of the reach and the whole lotic (and lentic) communities appear to be relatively
stream. It is also at this point that empirical studies persistent despite large seasonal fluctuations in larval
first begin to overlap with mathematical models of numbers (e.g. Townsend, Hildrew & Schofield, 1987;
food webs, which have driven theoretical develop- Weatherley & Ormerod, 1990; Woodward et al., in
ments and have been constructed to examine inter- press).
generational population dynamics (e.g. May, 1972, Geology, hydrology, land-use and habitat frag-
1973; Pimm, 1980). mentation can have strong influences upon the
At the scale of the whole stream and above, species-pool at the landscape scale and trophic
intergenerational dynamics and dispersal become interactions among the members of this pool form
increasingly important (e.g. Palmer, Allan & Butman, the community food web (Allan & Johnson, 1997;
1996; Hughes et al., 1998), and the current lack of Ward, 1997; Belyea & Lancaster, 1999). Energy and
understanding of the adult stage of many aquatic nutrients flow through this network of feeding links,
insects that disperse across the landscape, is obvious. influencing ecosystem processes, such as primary
Historically, freshwater ecologists have focused production and decomposition. As all food webs are
almost exclusively upon the larval insects that imbedded within, and constrained by, the surround-
dominate the benthos (Peckarsky & McIntosh, ing landscape, spatial position has important conse-
1998). Very little is known of the role of the adult quences for web pattern and process (Polis et al.,
in the population dynamics of riverine insects (but 1997). This is particularly true for lotic systems,
see Peckarsky et al., 1993; McPeek & Peckarsky, 1998; because they are closely coupled with the neigh-
Speirs et al., 2000). The few data that are available on bouring terrestrial environment (Hynes, 1975) that
adult dispersal provide apparently conflicting supplies basal resources to the food web (Hildrew,
results, depending upon the methods used. Conven- 1992; Tavares-Cromar & Williams, 1997; Wallace
tional ecological research, carried out with a variety et al., 1999) and habitat for adult aquatic insects
of traps, suggests that most adults stay close to their (Petersen et al., 1999). In addition to discrete patchi-
natal stream (e.g. Petersen et al., 1999), whereas ness within riverine systems (e.g. catchments con-
population genetics approaches often suggest that taining intermittent blocks of forestry interspersed
populations are well-mixed at large scales (e.g. with agricultural areas), to which we will return
Ó 2002 Blackwell Science Ltd, Freshwater Biology, 47, 777–798Riverine food webs 781
later, food web structure (e.g. predator–prey ratios)
Discontinuities along the river corridor:
and ecosystem functions (e.g. energy flow) can vary
geology, habitat fragmentation and land use
progressively along the river corridor (e.g. Habdija,
Radanovic & Primc-Habdija, 1997; Rosi-Marshall & In addition to continuous and gradual shifts in food
Wallace, 2002). web structure, discrete changes also occur along the
river corridor. For instance, the interrelationships
between acidification, geology and land-use can pro-
Source to sink: longitudinal gradients in riverine webs
duce large-scale discontinuities in the form and
Two well-known and much-debated conceptual mod- functioning of riverine food webs. Acidification of
els, the River Continuum Concept (RCC; Vannote fresh waters affects large areas of Europe and North
et al., 1980) and the Serial Discontinuity Concept America, particularly where deposition of atmo-
(Ward & Stanford, 1983) have been applied frequently spheric pollutants is high and the underlying geology
to explain longitudinal changes in natural and regu- of the catchment results in poorly buffered surface
lated rivers, respectively, although not always suc- waters: in such areas, acid-sensitive species may be
cessfully (see Winterbourn, Rounick & Cowie, 1981). lost from the riverine landscape, as pH falls (Hildrew
Examples of documented longitudinal changes in & Ormerod, 1995). Rainfall and geology also influence
food web structure include shifts in the dominant soil characteristics, which, in turn, determine the
resource base (e.g. Webster & Meyer, 1997) and the dominant land-use: this has further impacts upon the
structure of feeding guilds within the web, including riverine food web. Coniferous plantations in upland
predators (e.g. Habdija et al., 1997), filter-feeders (e.g. areas of the United Kingdom, for example, are often
Voelz & Ward, 1996) and herbivores (e.g. Hefti & found in discrete but large blocks on unproductive,
Tomka, 1991). Nutrient spiralling occurs as resources marginal land that is also geologically sensitive to
are assimilated by organisms and released progres- acidification (Rutt, Weatherley & Ormerod, 1989).
sively downstream. The longitudinal availability of These plantations can exacerbate the effects of acidi-
basal resources, especially fine particulate organic fication [e.g. by increasing occult deposition and
matter (FPOM), is therefore determined by both mobilisation of aluminium within the soil (Hildrew
physical (i.e. flow) and biological (e.g. shredding of & Ormerod, 1995; Friberg, Rebsdorf & Larsen, 1998)].
leaf litter) processes (Fisher et al., 1982; Naiman et al., During acidification the ‘traditional’ grazer assem-
1987; Crowl et al., 2001). blage is often lost or degraded (Ormerod et al., 1987)
A recent attempt to link food web theory more with generalist detritivore/herbivores becoming the
explicitly to the RCC (Rosi-Marshall & Wallace, 2002), dominant primary consumers (e.g. Ledger & Hildrew,
by studying energy flow through a macroinvertebrate 2000a,b). Conifer plantations can suppress the relative
food web, has shown that, although food web struc- importance of primary production within streams
ture did not vary markedly, the flow of organic matter even further, via increased shading and elevated
through the food web increased downstream. Along inputs of terrestrial detritus, resulting in food webs
the river corridor (> 30 km) the flux of organic matter that are driven by allochthonous subsidies of basal
through the food web varied by an order of magni- resources (Hildrew, 1992). The consequent shifts in
tude, and exceeded 1 kg m–2 year–1 in the down- the resource base from autotrophy towards heterotro-
stream reaches. The dominant food source switched phy may lead to greater donor-control (e.g. Hildrew,
from coarse particulate organic matter (CPOM), in the 1992; Dobson & Hildrew, 1992) which could increase
form of leaf litter (58% of consumption in the web stability [i.e. reduced vulnerability to species
headwaters), to suspended FPOM (64% of consump- extinctions and/or invasions (De Angelis, 1975;
tion in the downstream reaches) from the upper to the Pimm, 1982)]. Higher in the food web, the loss of fish
lower reaches, and algal and animal consumption as a result of reduced pH (Brown & Sadler, 1989)
increased fivefold and ninefold, respectively. The link potentially results in ‘mesopredator release’ (sensu
between resource availability and energy flow con- Courchamp, Langlais & Sugihara, 1999) of large
curred with the broad predictions of the RCC, but invertebrates, which can become extremely abundant
further studies of this kind are needed in a range of (e.g. Hildrew et al., 1984) (Fig. 2). Typically, these
systems. invertebrate predators are trophic generalists,
Ó 2002 Blackwell Science Ltd, Freshwater Biology, 47, 777–798782 G. Woodward and A.G. Hildrew
age, channelisation), the use of agrochemicals has
more diffuse, although powerful, effects upon food
web structure (Delong & Brusven, 1992). For example,
because organochlorine pesticides accumulate in the
body tissue of consumers they become increasingly
concentrated as they pass up through the food web
and can reach extremely high concentrations in the
top predators (Kidd et al., 1995). These chemicals have
been implicated in the impaired reproduction, or even
the extirpation, of many riverine species across large
areas, including snapping turtles, Chelydra serpentina
(Linnaeus) (de Solla et al., 2001), and otters, Lutra lutra
(L.) (Mason, 1991).
Eutrophication, through the addition of nitrate and
Fig. 2 Hypothetical interaction webs in circumneutral and phosphate fertilisers, can have very damaging effects
acidified freshwaters (adapted from Hildrew, 1992). Narrow upon riverine webs (Vandijk et al., 1994; Biggs, 1995).
solid lines represent weak trophic interactions; thick lines Algal blooms and the associated anoxia during die-
represent strong interactions with the direction of the main
effect indicated by arrows; horizontal broken lines represent
back have altered many riverine (and lentic) food
incidences of ‘self-damping’ (e.g. intraspecific, density-depend- webs dramatically. Eutrophication affects the physical
ent competition or cannibalism). Size of circles crudely repre- structure of the ecosystem (smothering substrata;
sents the abundance of trophic elements. The role of algae and inhibiting light penetration) and also the biological
their grazers in acid streams remains uncertain, although recent
composition of the community [e.g. an increase in
evidence suggests that algae are grazed by trophic generalists
(Ledger & Hildrew, 2000a,b). chironomids and oligochaetes; a decrease in oxygen-
sensitive taxa, such as stoneflies; altered periphyton
possibly because of the low productivity of acid production (Mason, 1991)]. Also, by altering the
streams (Hildrew et al., 1985). This dietary generalism C : N : P ratio of the resource base (e.g. Ebise &
will increase the reticulation of the food web, which Inoue, 1991), eutrophication may have drastic conse-
may also enhance its stability, as diffuse links can quences for the stoichiometry (sensu Elser et al., 2000)
weaken the ability of predators to induce trophic of the food web (see below).
cascades (e.g. Polis, 1991). The clearance of riparian vegetation, which often
Habitat fragmentation, both in the surrounding goes hand-in-hand with agricultural intensification,
terrestrial habitat and within the river channel itself, can deny the adults of many aquatic insects a suitable
can alter riverine food webs via effects upon dispersal, habitat within which to mate and oviposit (e.g.
metapopulation dynamics and the attendant, larger Petersen et al., 1999) and can remove a major energy
scale effects related to island biogeography and gene source in the form of allochthonous detritus (e.g.
flow. For example, the recovery of acidified streams in Dobson & Hildrew, 1992; Hall, Wallace & Eggert,
Wales following catchment liming has not followed 2000). In addition, the terrestrial arthropods that fall
the trajectory that was predicted, with some taxa from the riparian zone into the water provide a
showing little or no signs of returning to preacidifi- potentially important energy subsidy to the aquatic
cation levels. It has been suggested that the distances food web (e.g. Townsend & Hildrew, 1979; Mason &
from suitable sources of colonists were too large for Macdonald, 1982; Cloe & Garman, 1996; Nakano,
successful recolonisation (Rundle, Weatherley & Orm- Miyasaka & Kuhara, 1999; Kawaguchi & Nakano,
erod, 1995), although these ideas have been ques- 2001). Thus, management of the terrestrial landscape
tioned recently (Bradley & Ormerod, 2002). has the potential to alter the riverine food web
Agriculture has the most profound influence upon dramatically.
riverine webs at the landscape scale. These effects are Agricultural development also often alters the
often the converse of those associated with coniferous physical and hydrological characteristics of the river-
afforestation. In addition to the physical changes ine system. Drainage, sediment removal and chan-
associated with agriculture (e.g. deforestation, drain- nelisation have been the typical approach to the
Ó 2002 Blackwell Science Ltd, Freshwater Biology, 47, 777–798Riverine food webs 783
management of river corridors in both agricultural
Subsidies to and from riverine food webs: links forged
and urban areas (Allan, 1995). The increased flashi-
across system boundaries
ness, combined with the reduced availability of
bedform structures and storage zones that provide Moving further up the spatiotemporal scale, the
flow refugia, can potentially exclude certain species importance of food web linkages across habitat
from the food web and alter the strength of feeding boundaries and between ecosystems is apparent.
links (e.g. Wootton, Parker & Power, 1996; Chase, Exchange across ecosystems can have dramatic effects
2000). upon the food web (e.g. Polis et al., 1997; Nakano
et al., 1999) and riverine webs are no exception, being
open to exchange with marine (Humborg et al., 1997;
Lateral gradients: the flood plain and terrestrialisation
Milner et al., 2000), terrestrial (Wallace et al., 1997;
of riverine food webs
Kawaguchi & Nakano, 2001) and lentic (Wotton et al.,
In addition to any vertical and longitudinal gradients 1998) ecosystems. In the former, nutrients may be lost
in food web structure from the river’s source to its via export to the estuarine food web and, ultimately,
mouth, the lateral component of the river increases to the ocean. Catadromous fish, for example, spend
downstream as the flood plain broadens. This blurs most of their lives in freshwaters, only returning to the
the boundaries of the aquatic-terrestrial ecotone (Salo, sea to breed (e.g. eels, Anguilla spp.). Anadromous
1990) and provides an important habitat for many fish, in contrast, can transport considerable amounts
wetland species, particularly birds (Tomovcik, 1999). of potentially limiting nutrients (e.g. N in protein;
Migratory wildfowl and waders that feed on the flood marine-derived sulphur) and pollutants from the sea
plain not only link the terrestrial and aquatic systems to the riverine landscape (Ewald et al., 1998; Milner
laterally, but they can also link many different riverine et al., 2000). Such external inputs can be important in
food webs and transport nutrients and potential determining key phases of community succession
animal or plant colonists over thousands of kilometres across the landscape. For example, the first appear-
(Polis et al., 1997). Wetlands, like estuaries at the ance of Pacific salmon (Oncorhynchus spp.) following
longitudinal extremity of the riverine food web, are glacier retreat in Alaska is soon followed by an
among the most productive ecosystems on earth, and increase in bear populations, which feed extensively
vast energy and nutrient fluxes pass through them upon the migrating fish (Hilderbrand et al., 1999).
(Bayley, 1995). The flood plain is particularly import- Salmon carcasses that accumulate in the stream after
ant in tropical food webs (e.g. Winemiller, 1990, 1996), spawning can increase productivity by subsidising
but has been severely degraded in many temperate the macroinvertebrate community (Wipfli, Hudson &
regions, as a result of channelisation and drainage for Caouette, 1998; Wipfli, Chaloner & Caouette, 1999).
human habitation or agriculture (Ward, 1997; Ward There is also evidence that marine-derived nitrogen,
et al., 1998). Terrestrialisation of the riverine food web transported upstream via migrating salmon, becomes
increases with lateral distance from the channel, incorporated into the tissues of the streamside ripar-
resulting in the creation of more temporary, isolated ian vegetation, which ultimately re-enters the stream
lentic habitats further into the floodplain (Winemiller, as allochthonous detritus (Hilderbrand et al., 1999;
1990; Ward et al., 1998). Many of these habitats Milner et al., 2000). Thus, there can be important links
provide refugia for taxa such as amphibia, which and feedbacks among the marine, freshwater and
often suffer from fish predation in permanent water terrestrial food webs, despite the fact that these
bodies, but can be important consumers when fish are systems are usually considered as isolated, discrete
absent (e.g. Werner & McPeek, 1994; Wilbur, 1997). entities.
The hyporheic zone can also have an extensive lateral, External subsidies to riverine systems from terrest-
as well as vertical, component that stretches well rial systems (e.g. Flory & Milner, 1999; Kawaguchi &
beyond the confines of the river channel and may Nakano, 2001; Wipfli & Gregovich, in press) can
provide an important habitat for meiofauna and produce a range of direct and indirect food web
smaller macrofauna (Brunk & Gonser, 1997; Ward effects, including ‘apparent trophic cascades’ (e.g.
et al., 1998; Malard et al., 2002). Nakano et al., 1999). In systems with meagre food
Ó 2002 Blackwell Science Ltd, Freshwater Biology, 47, 777–798784 G. Woodward and A.G. Hildrew
resources, terrestrial invertebrates can provide a sub- organic matter derived from the lake form the basis of
stantial subsidy for the aquatic predators (e.g. Town- the filter-feeding guild of the riverine food web
send & Hildrew, 1979) that could, potentially, result in immediately downstream of the outflow (Robinson
a strengthening of apparent competition among prey & Minshall, 1990). Filter-feeding taxa, such as the
(sensu Holt, 1977) by sustaining predators at ‘artifici- Simuliidae, Hydropsychidae and members of the
ally’ high densities. Consequently, we might expect a Chironomidae (e.g. Rheotanytarsus spp.), can reach
level of contingency to exist in food web responses to very high densities and process phenomenal amounts
subsidies, which may vary with the degree of predator of food, removing suspended matter and transferring
specialisation, feeding behaviour (e.g. opportunistic it into the benthic food web (e.g. Malmqvist, Wotton &
versus switching behaviour) and/or the level of Zhang, 2001).
resource supply. Both mathematical models (e.g.
McCann et al., 1998) and empirical data (e.g. Wallace
et al., 1997) have demonstrated the importance of Quantifying riverine food webs
external inputs of consumers or resources to food
Connectance, energy flow and population dynamics
webs. In a large-scale experiment Hall et al. (2000)
found that per unit biomass consumption by preda- The traditional focus of food web ecology, particularly
tors was higher when terrrestrial detritus was exclu- in empirical studies, has been to list species and the
ded, suggesting increased interaction strength. The feeding links between them. This has resulted in a bias
magnitude of external subsidies and the feeding towards focusing upon patterns (e.g. connectance;
preferences of the consumers appear to influence the predator–prey ratios), rather than processes (e.g.
stability of the food web. Huxel & McCann (1998) energy flow, population dynamics) (Hall & Raffaelli,
found that, at low to medium levels of subsidy, 1993). Such qualitative, presence or absence, data also
modelled webs became more stable if species fed tend to overemphasise trivial interactions, because all
preferentially upon autochthonous resources, but species and links are given equal weighting (Paine,
unstable (and species lost) if the input was increased 1988; Benke & Wallace, 1997). The familiar connectance
and/or consumers favoured the allochthonous webs that dominated the early food web literature (e.g.
resources. Such cross-system subsidies can clearly Cohen, 1978; Pimm, 1982) are slowly being superseded
have far-reaching consequences for both energy flow by more quantitative data that provide more realistic
and population dynamics, which ecologists are only representations of natural food webs (e.g. Paine, 1992;
now starting to appreciate. For example, Kawaguchi & Raffaelli & Hall, 1996; Tavares-Cromar & Williams,
Nakano (2001) found that the form of riparian land 1996; Hall et al., 2000; Benke et al., 2001). These quan-
use affected the input of terrestrial prey, which titative descriptions have resulted in a dichotomous
determined the consumption by salmonids; this may approach to food web ecology, with one branch
alter the strength of ‘apparent trophic cascades’ within focusing upon population dynamics (e.g. Power,
the macroinvertebrate assemblage of the aquatic web 1990; Wootton et al., 1996), and the other upon energy
(e.g. Nakano et al., 1999). Thus land-use, ecosystem fluxes (e.g. Benke & Wallace, 1997), with virtually no
energy fluxes and the internal dynamics of the food studies that attempt to combine the two (but see
web can be intricately coupled in riverine landscapes. Hall et al., 2000). Even within these two branches
There exists a plethora of matter and energy fluxes, there is no standardisation as to how food webs are
other than the terrestrial–aquatic link, with which lotic quantified (Cohen et al., 1993). In particular, linkage
ecologists are probably most familiar, between rivers strengths are expressed in many ways (e.g. compare
and their surrounding ecosystems that have yet to be Tavares-Cromar & Williams, 1996; with Benke &
examined in detail within the conceptual framework Wallace, 1997), thus preventing any meaningful
of broader food web theory (Polis et al., 1997). For meta-analyses. The need to standardise units of meas-
example, lotic–lentic exchange of nutrients occurs at urement, especially when comparing across systems,
lake inflows and outflows, but the ramifications of has been emphasised repeatedly but has yet to be
these fluxes for the larger food web beyond these resolved (Cohen et al., 1993). By expressing carbon flux
relatively small ‘oases’ of secondary production are in terms of g C m–2 year–1 we have a baseline
not well understood. Plankton and dead particulate for characterising energy flow and, similarly, by
Ó 2002 Blackwell Science Ltd, Freshwater Biology, 47, 777–798Riverine food webs 785
measuring per capita consumption of individuals has recently been questioned (Lancaster & Waldron,
m–2 year–1, we can standardise for those studies that 2001) and fractionation appears to be far more
examine population dynamics, whilst also approxima- variable among species and habitats than is generally
ting to the units used in most models (Cohen et al., 1993). assumed (France, 1996; Finlay, Power & Cabana,
1999). In addition to these sampling problems, current
dual isotope models based on Euclidean distances
Modes of quantification: gut contents
between predator and prey overestimate the relative
and stable isotopes
importance of rare prey (Ben-David et al., 1997). This
Detailed gut contents analysis (GCA) provides evi- mathematical problem has, to date, hindered the use
dence of per capita consumption rates, and can be used of SIA in complex, multispecies food webs and
to assess the impact of predators upon prey popula- producing more accurate mixing models is now an
tions (e.g. Hildrew & Townsend, 1982; Speirs et al., important focus within this field (e.g. Phillips, 2001;
2000). Gut contents analysis can also be used to Phillips & Gregg, 2001; Ben-David, 2001). By combi-
estimate energy fluxes (e.g. Benke & Wallace, 1997; ning GCA with SIA, however, to produce a more
Hall et al., 2000). However, GCA is restricted in integrated approach, riverine ecologists should be
usefulness because it provides only a snapshot of a able to characterise trophic interactions in greater
predator’s diet, which can be extremely temporally detail and with greater accuracy than has been
variable. Results from GCA are also susceptible to achieved before (e.g. Hall et al., 2000).
errors caused by low sampling effort, as many
predators have no recognisable prey in their guts;
Integrating population dynamics and energy flow:
several hundred individuals may need to be exam-
food web stoichiometry
ined to describe the feeding links of a single predatory
species (Woodward & Hildrew, 2001). Also, the The next challenge is to produce a holistic view of
importance of food sources that are relatively amor- riverine food webs that links the currently disparate
phous (e.g. filamentous algae, soft-bodied prey, bio- fields of energetics and population dynamics (Fig. 3).
film, detritus) can be difficult to quantify using GCA. Combining GCA and SIA is a step in this direction,
An alternative, and increasingly popular, technique but we can take the process further. Food webs that
that overcomes some of these problems is the use of simply describe changes in population parameters
stable isotope analysis (SIA) (e.g. Cabana & Rasmus- without being constrained explicitly by thermody-
sen, 1996; Hall et al., 2000; Finlay, 2001). This approach namics or chemistry (i.e. via mass balance of ele-
provides an integrated measure of assimilation, and ments) may overlook important patterns in energy
allows ecologists to assess the relative contributions of flow, particularly where nutrient imbalances occur
resources with distinct isotope signatures to the food between consumers and resources (Sterner, 1995;
web (Lajtha & Michener, 1994). It can also provide Elser & Urabe, 1999). Conversely, the contribution of
information on ‘trophic status’, which is a continuous, individual taxa to the flux of energy or matter is often
rather than discrete, variable in systems with preval- obscured by lumping different populations into
ent omnivory, as is probably true for most real food ‘trophic species’ (sensu Cohen, 1978). Individual
webs (Polis, 1998; Williams & Martinez, 2000; Benke members within a guild can differ widely in their
et al., 2001). Essentially the differential enrichment of chemical composition, and this has implications for
carbon isotopes, expressed as d13C (&), reveals the food web processes. For example, shifts in species
contributions of different food sources to the food web composition can alter biogeochemical cycles, and
and enrichment of nitrogen, expressed as d15N (&), nutrient imbalance within a consumer’s diet can affect
reveals details of the trophic position. Because these not only its own growth rate but also that of its
isotopes appear to undergo predictable fractionation predators and prey (Elser & Urabe, 1999).
(‘you-are-what-you-eat-plus-one’ for d13C and ‘you- Recent attempts to unite the two disciplines have
are-what-you-eat-plus-three’ for d15N), biplots of the led to the development of ‘trophochemical’ webs
two signatures have been used to provide a simple (Fig. 4), which incorporate data on nutrient pools and
means of representing food webs quantitatively. trophic dynamics (Sterner, 1995). Imbalances in
However, the statistical rigour of many such studies nutrient flows (particularly shifts in C : N : P ratios)
Ó 2002 Blackwell Science Ltd, Freshwater Biology, 47, 777–798786 G. Woodward and A.G. Hildrew
Fig. 4 A trophochemical diagram of a hypothetical web
containing six species (after Sterner, 1995). For each species the x
and y coordinates give the quantities of elements X (e.g. C) and Y
(e.g. N), respectively, and the circular area denotes the quantity
of element Z (e.g. P). Solid arrows indicate trophic links (i.e.
species 6 consumes species 1, etc.) The dashed line represents the
1 : 1 Y : X ratio. The angle between the dashed and dotted lines
i
CD represents the angular imbalance between species 1 and 2.
1999; for a more detailed review): because any given
species must keep the C : N : P ratio in its body
Fig. 3 Schematic representations of the three major approaches tissues within certain narrow limits, changes in the
to the study of food webs. In the connectance web all species are C : N : P ratio among food resources determine which
of equal importance (i.e. the size of the nodes are identical) and
consumers will be most successful and, in turn,
feeding links are also weighted equally (i.e. line widths are
identical), because the web is constructed from presence/ determine the species composition within the web.
absence data. The two quantified webs show that the web Nutrient imbalances in primary consumers can influ-
structure is skewed among species and links. However, the ence which predators occupy the higher trophic levels
species that contribute most to the energy flow in the web do not and also which primary producers are most success-
necessarily have an appreciable effect upon the population
dynamics of the web. For example, large species that consume
ful. The greatest progress to date has been made in the
considerable amounts of biomass and dominate secondary pelagic zone of lakes, where preliminary results
production may be numerically rare, whereas small species may suggest that stoichiometric imbalances in C : N : P
be very common in the benthos and in predator diets, but have ratios can have profound consequences for plankton
little effect upon energy flow.
population dynamics and may produce alternative
stable states (Elser & Urabe, 1999). However, such
through these webs have led to the emergence of the techniques are rarely applied in the study of lotic food
relatively new field of food web stoichiometry, which webs, and no studies have yet been published that
links energy flow and population dynamics to pro- include the benthos (Elser & Urabe, 1999; Elser et al.,
duce a more integrated approach to understanding 2000). Although there are plenty of studies that
food webs (Huxel & McCann, 1998; Polis, 1999). The examine N : P or C : N ratios, the three-way dynam-
basic idea of applying stoichiometry to food web ics among these elements that appear to be important
theory is summarised as follows (see Elser & Urabe, in moulding webs, are still largely unknown.
Ó 2002 Blackwell Science Ltd, Freshwater Biology, 47, 777–798Riverine food webs 787
Recent research on nutritional constraints in aquatic and/or species richness) enhances stability, if most
food webs suggests that highly unsaturated fatty acids links are weak. The distribution of linkage strengths is
(HUFA), which are synthesised almost exclusively by also important, with stability declining as web struc-
plants, are required by many grazers to maintain high ture becomes more skewed; as a few links increase
growth, reproduction and assimilation rates (Brett & their influence upon web dynamics the prevalence of
Müller-Navarra, 1997). Because HUFA availability trophic cascades and species extinctions rises (Borrv-
varies among plant species, spatial changes in the all, Ebenmann & Jonsson, 2000). Thus, the more
species composition of the primary producers can diffuse interactions are, the less likely the web is to
have strong effects upon the food web by altering shift in response to a perturbation. These findings
food quality and, consequently, grazer abundance hark back to some of the earliest ecological ideas
and the strength of feeding links. Although such about community and food web structure (e.g. Elton,
studies have, like stoichiometric analysis, been largely 1927, 1958; MacArthur, 1955).
restricted to lentic systems, similar nutritional con- More stable systems should be less prone to trophic
straints may be important in rivers, where primary cascades and switching between ‘alternative stable
producer and grazer assemblages can vary dramatic- states’. Although it was once thought that trophic
ally over large spatial scales. cascades were ‘all wet’ (Strong, 1992), being special
Large scale switches, along the river corridor, from cases of aquatic food webs, they have recently been
terrestrial detritus (very C enriched) to an algal/ reported in a diverse range of ecosystems (Pace et al.,
aquatic macrophyte resource base should, theoretic- 1999; but see Polis et al., 2000). However, although
ally, have strong effects upon the C : N : P ratios of cascades certainly occur in some riverine food webs,
the secondary producers within the web, the conse- they appear to be far from ubiquitous. For example,
quences of which remain unexplored. A downstream Flecker & Townsend (1994) and Biggs et al. (2000)
decline in the allochthonous input of C may even lead have reported algal blooms as a result of the (indirect)
to C limitation of algal uptake in streams with very effects of introduced trout in New Zealand streams.
large catchments (e.g. Finlay, 2001). If the stoichio- Similarly, Power (1990) and Wootton et al. (1996) have
metric patterns seen in standing waters [and terrest- also described strong cascading interactions. On the
rial systems (Elser et al., 2000)] apply to riverine other hand, however, Woodward & Hildrew (2001)
systems, changes in C : N : P ratios could have found relatively little change in prey abundance
profound implications for web structure as produc- within the complex (i.e. highly interconnected) food
tivity increases (e.g. via eutrophication), the effects of web of a headwater stream, following the invasion of
which may reach across ecosystem boundaries (e.g. a new top predator. It would seem that certain
Ebise & Inoue, 1991; Justic, Rabalais & Turner, 1995; characteristics of particular webs and species deter-
Humborg et al., 1997). mine whether strong effects, such as cascades, will
occur. The challenge is now for freshwater ecologists
to identify those systems and species most vulnerable
From pattern to process: trophic cascades
to disruption.
and trickles in riverine webs
We have several suggestions that could serve to
Food web ecology has undergone a fundamental identify vulnerable webs. If recent theoretical and
‘paradigm shift’ in recent years, with the established empirical advances are borne out, then we could
view that simplicity begets stability (e.g. May, 1972, expect that more linear webs, with more discrete
1973; Pimm, 1982) being apparently overturned (Polis, ‘trophic levels’, will be more prone to cascades and
1998). Freshwater systems have played an increas- species extinctions than short, broad and intercon-
ingly pivotal role in changing our perceptions of food nected generalist webs (e.g. Borrvall et al., 2000).
webs. For example, three of the six sources used to Similarly, generalist predators may cause less disrup-
develop the recent ‘niche model’ of food web struc- tion than specialist predators, although if the gener-
ture of Williams & Martinez (2000) were from fresh alists do not exhibit prey switching apparent
waters. Essentially, the results of recent advances in competition may result in the loss of some of the
mathematical modelling and field experiments now rarer prey. Understanding how webs are structured,
suggest that complexity (i.e. high linkage density and which sorts of interactions might be important,
Ó 2002 Blackwell Science Ltd, Freshwater Biology, 47, 777–798788 G. Woodward and A.G. Hildrew
provides the key to the first steps in predicting headwater streams (e.g. Schmid-Araya & Schmid,
changes in food web structure in response to pertur- 2000; Woodward & Hildrew, 2001; Schmid-Araya
bation. The suggested importance of nutrient supply, et al., in press). Consequently, trophic cascades might
top–down control and stoichiometry for the mainten- well be weaker in some riverine webs because of
ance of alternative stable states within shallow lakes trophic generalism and high omnivory, which appear
(Scheffer, 1998; Elser & Urabe, 1999), via trophic to be particularly prevalent in headwaters and/or less
cascades, may well also be applicable to some riverine productive systems (e.g. Schmid-Araya & Schmid,
systems, at least where similar cascading patterns 2000; Woodward & Hildrew, 2001). Thus, the stability
arise (e.g. Power, 1990). of a web, in terms of its vulnerability to cascades, may
In addition to deterministic processes acting in be contingent upon its spatial positioning within the
isolation, stochastic disturbance can affect cascades, landscape. The assertions of Chase (2000) therefore
via its effect upon these processes. Spates can exacer- need to be examined and tested more rigorously in a
bate trophic cascades in large rivers, by removing diversity of riverine systems, if they are to be widely
species that are ‘more defended’ (i.e. k-selected and accepted as typical of lotic webs.
less vulnerable to predators), thereby increasing the
strength of interactions between predators and the
Macroecology of riverine webs: beyond
remaining, more vulnerable, r-selected prey (Wootton
community ecology
et al., 1996). Cascades also tend to increase with
nutrient enrichment, but can be stabilised by spatial Recent papers have stimulated interest in exploring
heterogeneity and refugia (Pace et al., 1999). In addi- relationships between species abundance and bio-
tion, trophic generalism and omnivory within a food mass, spatial scale and food web structure (e.g. Leaper
web can potentially weaken the strength of cascades & Raffaelli, 1999; Holt et al., 1999). Such studies are
(Hildrew, 1992; Pace et al., 1999). Chase (2000) has often bracketed within the realm of ‘macroecology’
argued that because disturbance in flowing waters is (sensu Brown, 1995), which is ‘a blend of ecology,
far greater than in standing waters, this constrains the biogeography and evolution’ (Lawton, 1999, p. 182),
ability of species to exploit defensive (i.e. antipreda- emphasising the large temporal and spatial scales
tor) traits. Consequently, lotic systems should have involved (e.g. Brown & Maurer, 1989). Macroecology
less defended species in simpler food chains and may be considered to occupy a tier above community
therefore may be more prone to trophic cascades than ecology, a field which Lawton (1999) suggests is so
lentic systems (Chase, 2000). bedevilled with contingency that the validity of
Spates, however, do not necessarily constitute searching for generalities is questionable (but see
disturbance per se. The recovery of stream communi- Williams & Martinez, 2000 for an example of a simple
ties following high flows can be extremely rapid, and successful descriptor of community food web
because of the availability of in-stream flow refugia, structure).
storage zones and the hyporheos (e.g. Lancaster &
Hildrew, 1993; Ward et al., 1998). Further, riverine
Predator : prey ratios, species–area relationships
species can also use an abundance of spatial refugia
and food web structure
from predators within the structurally complex ben-
thos, thereby weakening or decoupling predator–prey An example of a macroecological pattern that is
interactions (Williams, Barnes & Beach, 1993; Hillb- particularly pertinent to riverine food webs is the
richt-Ilkowska, 1999). Thus the morphology of the posited constant ratio of predator to prey species in
stream and its surrounding landscape can have fresh waters, reported by Jeffries & Lawton (1985) as
important consequences for the food web. an example of large-scale trophic structuring across
Anti-predator traits are well known in many lotic systems. Apparent competition (sensu Holt, 1977) has
species [e.g. crypsis, avoidance responses and chem- been suggested as a process that might produce this
ical defences (Allan, 1995)]. Also, the evidence that pattern (Mithen & Lawton, 1986), although several
lotic food webs are simpler than lake food webs is other mechanisms that operate in a hierarchical
equivocal (see Bengtsson, 1994) and more recent data manner over a range of spatiotemporal scales may
suggest that the opposite may be true, at least for also create the same effect (Warren & Gaston, 1992).
Ó 2002 Blackwell Science Ltd, Freshwater Biology, 47, 777–798Riverine food webs 789
Holt et al. (1999), however, have suggested recently speculative, they have the potential to provide a
that the slope of species–area relationship, which they conceptual framework within which food web and
refer to as ‘one of the most robust empirical generalisations landscape ecology can be integrated.
in ecology’ (Holt et al., 1999, p. 1495) may vary with
trophic rank, being higher for predators than primary
Compartmentalisation of riverine food webs:
consumers. If this is true, then the assertion of a
do spatial, temporal and trophic subwebs exist?
constant predator : prey ratio may be wrong, as the
relationship should become curvilinear when plotted Another aspect of food web ecology that is pertinent at
against area. These ideas, which are still largely the landscape scale, but has been largely ignored,
speculative ((Holt et al., 1999) supply little empirical relates to the supposed positive relationship between
data to support their suggestions), could provide compartmentalisation and food web stability (e.g. May,
important insight into how riverine webs are struc- 1972, 1973; Pimm, 1982). Compartmentalisation refers
tured within the landscape. For example, do head- to the existence of relatively isolated ‘subwebs’ or
waters contain a lower proportion of predators than ‘blocks’ (sensu May, 1972, p. 414) within the larger food
large tributaries, which generally possess more spe- web of the entire community. Compartmentalisation
cies (e.g. Horwitz, 1978)? If so, does the prevalence of does not necessarily relate solely to spatial segregation
apparent competition among prey vary with stream (although it can), but can also arise where distinct
order (i.e. area)? If there is an area-dependent increase trophic groups (i.e. guilds) exist in the same space. For
in apparent competition does this stabilise larger example, although spatially overlapping, herbivores
riverine food webs, as predicted by some models (e.g. and detritivores may occupy relatively separate sub-
McCann et al., 1998)? It has been suggested that webs in some systems (e.g. Polis & Hurd, 1996), with
apparent competition can be particularly stabilising few strong interconnections between them. We might
when combined with intraguild predation (e.g. McC- expect the degree of spatial compartmentalisation to
ann et al., 1998), the likelihood of which should increase from the source to the mouth of a river, as
increase with the proportion of predatory species. habitat heterogeneity increases (e.g. large macrophyte
Thus, if the suggestions of Holt et al. (1999) and recent stands, gravel bars, open water, backwaters). This
food web models are correct, we might expect large localised spatial diversity should also lead to an
rivers to have far more stable webs, which are less increase in the number of species within the ‘regional’
prone to cascades and species extinctions, than head- food web (after Thienemann’s, 1954 First Law) (Hillb-
waters. richt-Ilkowska, 1999). In addition, feeding guilds
There are some tentative suggestions, however, that should diversify as detritivores are joined by grazers
riverine food webs might function somewhat differ- (Vannote et al., 1980), vertebrate predators join inver-
ently from these predictions, because of ‘peculiarities’ tebrate predators (Hildrew et al., 1984) and the benthos
in their structure. For example, an abundant guild of is augmented by pelagic taxa (Reynolds, Carling &
large invertebrate predators tends to characterise Beven, 1991). Although compartmentalisation has been
headwaters, as these species are often excluded from demonstrated in a range of systems, it was not detected
larger rivers by fish (e.g. Hildrew et al., 1984). Food in a well-described estuarine food web (Raffaelli &
webs in headwaters can have extremely high linkage Hall, 1992), and has not yet been examined systemat-
complexity, with the generalist predators preying on a ically and in detail within a truly riverine context.
large proportion of the prey assemblage (e.g. Schmid-
Araya & Schmid, 2000; Woodward & Hildrew, 2001).
Body-size, constraint space and web structure
This might suggest that apparent competition is
actually stronger in these smaller webs. However, it The body-size constraint space (sensu Brown, 1995),
is theoretically possible that the generalist nature of which bounds the relationship between body size and
these webs may weaken or reverse the effect of abundance within a community, has been shown to
trophic rank upon the species–area relationship, an hold true (at least as an approximation) for an
exception that is mentioned by Holt et al. (1999), and estuarine food web (Leaper & Raffaelli, 1999), but
this may explain the apparently contradictory nature has not yet been examined in food webs further
of lotic webs. Although these ideas are still largely upstream. We might predict that as we move
Ó 2002 Blackwell Science Ltd, Freshwater Biology, 47, 777–798You can also read
