Floodplain river food webs: Generalisations and implications for fisheries management

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ABSTRACT
Based on the relationship between temperature
variation and flood dynamics, three types of floodplain
rivers can be identified: temperate stochastic, temper-
ate seasonal and tropical seasonal. The degree to which
flooding occurs in phase with warm temperatures and
enhanced system productivity influences selection for
alternative life history strategies in aquatic organisms.
In addition, regional geochemistry and temporal
dynamics of disturbance and recovery of local habitats
within the landscape mosaic favour different life
history strategies, sources of production and feeding
pathways. In most habitats, algae seem to provide the
most important source of primary production entering
Winemiller K.O.
Department of Wildlife and Fisheries Sciences, Texas A&M University, College Station, TX, 77843-2258, United
States E-mail: k-winemiller@tamu.edu
FLOODPLAIN RIVER FOOD WEBS:
GENERALIZATIONS AND IMPLICATIONS
FOR FISHERIES MANAGEMENT
Keywords: connectivity, detritus, migration, primary
production, species interaction, trophic position
285

the grazer web. Large fractions of periphyton and
aquatic macrophyte production enter aquatic foodwebs
in the form of detritus and detrital consumption is
greater during low-water phases. Even in species-rich
tropical rivers, most of the material transfer in food
webs involves relatively few species and short food
chains (3-4 levels, 2-3 links). Longer food chains that
involve small or rare species are common and increase
ecological complexity, but probably have minor effects
on total primary and secondary production. In the trop-
ics, fishes appear to perform many ecological func-
tions performed by aquatic insects in temperate rivers.
Oftentimes, a small number of common species dis-
proportionately influences benthic ecosystem struc-
ture, productivity and dynamics. Similarly, a relatively
small number of predatory species may exert a dispro-
portionately large influence on prey populations, even
in species-rich tropical systems. Under seasonal flood-
pulse regimes, species have the opportunity to evolve
adaptations to exploit predictable resources. Under
aseasonal flood-pulse regimes, species are more chal-
lenged to respond appropriately to relatively unpre-
dictable patterns of resource variation and access to
floodplain habitats, while nonessential for most
species, usually enhances recruitment. Seasonal rivers
in nutrient-rich landscapes can sustain greater harvest
than aseasonal rivers or seasonal rivers in nutrient-
poor landscapes. Loss of habitat connectivity and over-
harvest of dominant species can have unpredictable
effects on food web dynamics and community struc-
ture. Maintenance of natural flood regimes is impor-
tant for biodiversity conservation and sustainable har-
vest of fishes, especially in strongly seasonal systems.
IMPORTANCE OF RIVER-FLOODPLAIN
SYSTEMS
River-floodplain systems, especially in the
tropics, support high biological diversity and important
fisheries (Welcomme 1985; 1990; Lowe-McConnell
1987). High biological diversity, both taxonomic and
functional, is associated with high spatial complexity
and the dynamic nature of aquatic, terrestrial and eco-
tonal habitats (Schiemer 1999; Ward, Tockner and
Schiemer 1999; Robinson, Tockner and Ward 2002).
River networks are ubiquitous features of landscapes
that have provided many opportunities for allopatric
speciation of aquatic taxa and also serve as reservoirs
that accumulate species over evolutionary time. The
high habitat heterogeneity and ecotonal nature of river-
floodplain landscapes also fosters high richness of ter-
restrial taxa.
The nutrient-rich alluvial soils often associated
with lowland floodplains have always been targets for
intensive agriculture. Use of floodplains for agriculture
has resulted in construction of levee systems to control
flooding. Levees sever aquatic connections between
the river channel and aquatic habitats of the floodplain
(Sparks 1995; Ward et al. 1999). In addition to direct
impacts from agriculture and other land uses that
destroy natural terrestrial, wetland and aquatic habi-
tats, lowland rivers are impacted by pollution, includ-
ing nutrient loading, from locations anywhere within
their catchments. The natural hydrology of most large
rivers in developed nations and increasingly in devel-
oping nations has been severely altered by dams, lev-
ees, channelization and landscape changes. In spite of
their great ecological, economic and cultural impor-
tance, large rivers remain one of the most poorly stud-
ied among major ecosystems (Thorp and Delong
1994). Recent years have witnessed an increase in
research on large rivers, especially in Europe, Australia
and the Americas. Even as we begin to understand the
ecology of large river ecosystems, with each passing
year fewer relatively un-impacted large rivers remain
as models for future restoration.
The purpose of this paper is to briefly review
food web structure and dynamics in lowland river-
floodplain systems and to explore management impli-
cations of this body of ecological knowledge. The food
web paradigm provides an approach that allows us to
model complex communities and ecosystems with the
286 Floodplain river food webs: Generalizations

ultimate aim of understanding relationships and pre-
dicting dynamics. The historic development of the
food web paradigm has been reviewed previously
(Hall and Raffaelli 1993; Polis and Winemiller 1996).
Woodward and Hildrew (2002) recently reviewed food
web structure in rivers, with a strong focus on theories
and evidence associated with system stability. Their
review emphasized evidence from streams, since com-
paratively little food web research has been conducted
on large rivers. The present review seeks to summarize
recent findings and perspectives from large lowland
rivers. Additionally, the features of lowland rivers from
tropical and temperate regions will be compared and
generalizations sought for application to conservation
of biodiversity, fisheries and ecosystem integrity and
productivity.
TYPOLOGY OF RIVER-FLOODPLAIN
ABIOTIC DYNAMICS
The degree to which flooding occurs in phase
with warm temperatures and enhanced system produc-
tivity influences selection for alternative life history
strategies in aquatic organisms. Rivers display at least
three general patterns: temperate with aseasonal
(seemingly random) flood pulses, temperate with sea-
sonal flood pulses and tropical with seasonal flood
pulses. The ramifications of these patterns for ecologi-
cal dynamics, food web dynamics in particular, are the
and implications for fisheries management 287
Figure 1. Examples of lowland floodplain rivers with temperate-aseasonal (Brazos River- from US Geologial Survey database), tem-
perate-seasonal (Illinois River- from Sparks 1995; Broken River- from Humphries et al. 2002) and tropical-seasonal (Zambezi River- from
Handlos and Williams 1985; Niger River- from Quensiere et al. 1994; Orinoco River- from Hamilton and Lewis 1990) abiotic regimes.
focus of this paper. Photoperiod and temperature are
key environmental drivers of ecological dynamics in
fluvial systems. Longer photoperiods during summer
support increased primary production. Warmer tem-
peratures increase rates of microbial metabolism,
nutrient cycling, primary production and feeding by
ectotherms. At high latitudes and elevations, spring
warming also is associated with snowmelt and
increased water availability. The effect of flooding on
feeding, growth and survival of aquatic organisms can
be particularly strong in lowland floodplain river sys-
tems. Floods stimulate remineralization of nutrients as
well as primary and secondary production in flood-
plain habitats (Welcomme 1985; Junk, Bayley and
Sparks 1989).
In temperate regions, temperature varies in a
predictable seasonal pattern, with the magnitude of
variation greater at higher latitudes and elevations.
Regions having fairly unpredictable rainfall and lack-
ing significant runoff from snowmelt display unpre-
dictable, aseasonal flood pulses. Examples of temper-
ate-aseasonal rivers are found along the northwestern
Gulf of Mexico coast of North America and in certain
regions within Australia’s Murray-Darling Basin. In
Texas, the Brazos River shows unpredictable hydrolo-
gy, both within and between-years (Winemiller 1996a,
Figure 1). High discharge events vary greatly in

magnitude and most are of short duration. Floods that
top riverbanks and enter oxbow lakes are infrequent
and can occur any time of the year (Winemiller et al.
2000). The unpredictable nature of flood pulses and
river-floodplain connections pose challenges for
species that exploit ephemeral or dynamic ecotonal
aquatic habitats.
Many temperate regions have cyclic patterns of
precipitation and/or springtime melting of ice and
snow that yield seasonal flood pulses. Local flooding
may derive from local precipitation and thawing (e.g.
Broken River, Australia; Illinois River, United States,
Figure 1), precipitation and/or snowmelt in headwater
areas (e.g. lower Colorado River, United States), or
some combination of local and upstream factors.
Seasonal flooding in the temperate rivers also can be
strongly influenced by evapotranspiration as a function
of seasonal temperature regimes (Benke al. 2000). The
magnitude of flooding in most temperate rivers is high-
ly variable between years (e.g. Ogeechee River, south
eastern United States, Benke et al. 2000) and in some
systems floods may not occur at all during some years
(e.g. Broken River, Australia, Humphries, Luciano and
King 2002). Thus, whereas temperate-seasonal rivers
provide a relatively predictable temporal regime to
which organisms may respond adaptively (Resh et al.
1994), stochastic between-year variation may serious-
ly challenge adaptive responses to seasonal environ-
mental periodicity. In most cases, seasonal flooding in
the temperate zone coincides with springtime warm-
ing, which selects for reproduction during this period.
Recruitment is enhanced when early life stages occur
in appropriate habitats when warm temperatures stim-
ulate ecosystem productivity, metabolism and growth.
In tropical continental regions, the flood pulse
of lowland rivers is almost universally driven by
strongly seasonal precipitation. In some cases, local
flooding coincides with local precipitation (Upper
Orinoco, Upper Paraná, Upper Zambezi and Fly
Rivers), whereas in others the seasonal flood pulse is
most strongly influenced by rainfall in distant headwa-
ters (e.g. lower Niger, Congo and Solimões-Amazon
Rivers). Because temperature varies relatively little in
tropical lowland regions, the hydrological regime is
the major factor that drives ecological dynamics and
natural selection in response to environmental varia-
tion. The tropical-seasonal model has dominated think-
ing about the ecology of river-floodplain systems (e.g.
the flood-pulse model, Junk et al. 1989), but global
generality of this pattern and its consequences has
scarcely been discussed (but see below, also Thorp and
Delong 1994, 2002; Humphries, King and Koehn
1999: Humphries et al. 2002).
PRIMARY PRODUCTION SOURCES FOR
LOWLAND RIVER FOOD WEBS
A fundamental aspect of any food web is the
source of primary production that supports consumer
populations. Geology and landscape features influence
nutrient and flood dynamics that affect production
rates of different primary producers (Rai and Hill
1984). Primary production has high spatiotemporal
variation within most river-floodplain systems. In the
central Amazon Basin, primary productivity ranges
from 50 to 3 500 mg C m-2 d-1 (Rai and Hill 1984)
according to location and flood stage. Macrophytes,
both terrestrial and aquatic, appear to be the major pro-
ducers in floodplains (Bayley 1989; Melack et al.
1999; Lewis et al. 2001). Analysis of stable isotopes
indicates that dominant production sources for higher
consumers in river-floodplain food webs appear to be
phytoplankton, periphyton and fine particulate organic
matter derived from algae (Araujo-Lima et al. 1986;
Hamilton, Lewis and Sippel 1992; Forsberg et al.
1993; Thorp and Delong 1994, 2002; Thorp et al.
1998; Benedito-Cecilio et al. 2000; Lewis et al. 2001;
Leite et al. 2002). Even in highly turbid floodplain
lakes of arid central Australia, benthic filamentous
algae in the shallow littoral zone are the major produc-
tion source supporting higher consumers (Bunn,
Davies and Winning 2003).
Both algae and aquatic macrophytes appear to
enter aquatic food webs mostly in the form of detritus
(fine and coarse particulate organic matter), some
being transported in the water column and some set-
tling onto substrates. Direct consumption of aquatic
macrophytes is rare, but aquatic macrophytes are con-
sumed by a few fish genera from South America
(Schizodon [Anosotomidae] and Pterodoras
[Doradidae]) and Africa (Tilapia [Cichlidae]).
Detritivory is extremely common in river communi-
ties, both among invertebrates and fishes. In seasonal
288 Floodplain river food webs: Generalizations

floodplain habitats of the Orinoco and Zambezi rivers,
consumption of detritus by fishes was greater during
low-water phases (Winemiller 1990, 1996a). As deter-
mined from analysis of stomach contents, fishes con-
sumed large fractions of both fine and coarse particu-
late material. In these systems, coarse detritus is
derived almost entirely from aquatic macrophytes. The
origin of fine particulate matter in diets could not be
determined from microscopic analysis, but isotopic
studies suggest mixtures of algae and macrophytes that
use the C3 photosynthetic pathway (Jepsen and
Winemiller 2002).
Based on isotopic evidence and the fact that
coarse particulate matter derived from macrophytes is
refractory and of poor nutritional value, Thorp and
Delong (1994, 2002) made a case for a dominant role
of algae in river food webs. In tropical-seasonal rivers,
macrophytes generally produce well over half of the
primary production on floodplains, yet only contribute
small fractions of the total carbon assimilated by fish-
es (Forsberg et al. 1993; Lewis et al. 2001).
Macrophyte production is high during the period of
floodplain inundation (Rai and Hill 1984; Welcomme
1985; Junk et al. 1989). As floodwaters recede, aquat-
ic macrophytes die and produce massive amounts of
coarse detritus, only a minor fraction of which is prob-
ably consumed in any form by aquatic macrofauna.
Most of the labile dissolved organic carbon leaches
from this material and is quickly consumed by
microbes. Most of the remaining refractory material
seems to be consumed by microbes (the microbial
loop), without direct entry into the upper food web
(Figure 2). The fraction of microbial carbon that makes
its way to the upper web is unknown for virtually all
rivers, but assumed to be small based on available iso-
topic evidence (e.g. Bunn et al. 2003). In eutrophic
floodplains, huge stocks of water hyacinths, grasses, or
other macrophytes build up during the flood phase. As
water levels drop, microbial metabolism of dead
macrophyte tissues can deplete dissolved oxygen with-
in shrinking aquatic habitats (Winemiller 1996b). In
many savanna floodplains, such as the Kafue flats of
the Zambezi system, submergence of terrestrial grass-
es during the rising-water phase leads to plant death,
decay and aquatic hypoxia over large areas (Junk et al.
1989).
and implications for fisheries management 289
Figure 2. Generalized food web for floodplain-river ecosystems. Boxes are aggregate material pools and vectors represent consumer-
resource interactions with thick arrows representing dominant pathways (ml= microbial loop path, fp = nutrient pathways enhanced by
flood pulses, iw = invertebrate web having complex trophic structure involving invertebrates and ? = poorly quantified pathways).

In tropical systems, terrestrial sources of pri-
mary and secondary production are directly consumed
by diverse fish taxa. In the central Amazon, several
abundant fish species consume seeds, fruits, arthro-
pods and other forms of allochthonous resources (e.g.
Goulding 1980; Goulding, Carvallo and Ferreira
1988). Some characiform fishes (e.g. Brycon,
Colossoma, Piaractus and Myleus spp.) are morpho-
logically and physiologically specialized to feed on
fruits and seeds. Goulding (1980) described large
amounts of fruit and seeds in diets of many Amazonian
catfishes (Siluriformes). Terrestrial invertebrates and
vertebrates also enter the aquatic food webs. The arua-
na (Osteoglossum bicirrhosum Spix and Agassiz) is
able to leap several meters above the water surface to
feed on arthropods, reptiles, birds and bats. Accounts
of direct consumption of allochthonous resources in
the flooded forests of the Amazon had a large influence
on the development of the flood pulse concept for large
rivers. Yet when the aquatic food web is viewed as a
whole (i.e. major biomass components) allochthonous
carbon sources appear to be less important for macro-
faunal populations than autochthonous sources of pri-
mary production. The greatest fraction of terrestrial
vegetation that enters river-floodplain food webs
appears to do so as detritus (leaf litter and woody
debris), most of which is highly refractory and
processed via the microbial loop.
FOOD WEB STRUCTURE
River food webs are extremely complex and
dynamic (Winemiller 1990). Yet one of the most strik-
ing features of river communities is the domination of
standing biomass by a relatively small number of
species. This pattern appears to be true both in low-
diversity temperate systems, but more surprisingly the
pattern holds also for taxonomically diverse biotic
assemblages in tropical rivers. Fishery yields from
almost every major floodplain-river system in the
world are strongly skewed in favour of a handful of
dominant species (e.g. see summaries in Welcomme
1985). In terms of standing biomass, the Orinoco and
Amazon river mainstems are dominated by a few
species of Prochilodus, Semaprochilodus, Mylossoma,
Hydrolycus, Brycon, Pseudoplatystoma, Pinirampus,
and Brachyplatystoma. Obviously, much biomass may
be represented by small fishes of little or no commer-
cial value, however, even these small fish assemblages
are strongly skewed with few abundant and many
uncommon species (e.g. Winemiller 1996b, Arrington
and Winemiller 2003). Thus, it is reasonable to assume
that matter and energy moving through a local food
web are doing so via a comparatively small subset of
the total pathways represented in the trophic network.
This was indeed the pattern demonstrated for the
aquatic food webs in four tropical freshwater systems,
including a creek-floodplain system in the Venezuelan
llanos and Atlantic coastal plain of Costa Rica
(Winemiller 1990). When the magnitude of trophic
links was estimated as the volumetric proportion of
resource categories in consumer diets, the distribution
of link magnitudes was strongly skewed in every
instance. In terms of biomass, relatively few dominant
producer and consumer taxa and a limited number of
major trophic pathways dominate river food webs.
Aquatic and terrestrial macrophytes usually are
dominant sources of primary production in floodplains
(Rai and Hill 1984) and most of this material is con-
sumed by microbes that ultimately return nutrients to
the inorganic pool (Figure 2). However, not all detritus
is recycled within the microbial loop, with variable
fractions consumed directly by a variety of inverte-
brate and fish taxa, some of which are dominant food
web elements. Important components of aquatic meio-
and macro-invertebrate faunas are detritivores
(Schmid-Araya and Schmid 2000; Benke et al. 1984;
Benke et al. 2001). Although the standing biomass of
these taxa is generally low, they have high rates of pop-
ulation growth and turnover and represent important
pathways in river food webs. Much more research is
needed to elucidate the functional significance of
aquatic invertebrates, particularly meiofauna, in large
river food webs.
Detritivorous fishes are always abundant in
river-floodplain systems and routinely dominate fish-
ery catches (Welcomme 1985). Although some detri-
tivorous fishes consume coarse vegetative detritus,
most of the material classified as detritus in gut
290 Floodplain river food webs: Generalizations

contents is fine amorphous material of undetermined
origin. Detritivorous fishes are important prey for large
piscivores. In the Cinaruco River of Venezuela,
Semaprochilodus kneri (Pellegrin) were estimated to
contribute about 45 percent of the diet of large Cichla
temensis Humboldt during the falling-water period
(Jepsen, Wimemiller and Taphorn 1997; Winemiller
and Jepsen 2002). Detritivorous fishes form major por-
tions of the diets of piscivorous catfishes in large South
American rivers (Barthem and Goulding 1997;
Barbarino and Winemiller unpublished). Tigerfish
(Hydrocynus vittatus Castelnau) and African pike
(Hepsetus odoe (Bloch) of the Upper Zambezi River
consume large numbers of detritivorous tilapines and
cyprinids, respectively. Yet isotopic evidence indicates
that comparatively little carbon from macrophytes,
especially grasses using the C4 photosynthetic path-
way, makes its way to higher consumers (Hamilton et
al. 1992; Lewis et al. 2001; Jepsen and Winemiller
2002). Information currently available from research
in large rivers in North and South America indicates
that much of the fine particulate organic matter assim-
ilated by detritivorous fishes is derived from algae,
even in systems in which aquatic macrophytes domi-
nate aquatic primary production (Araujo-Lima et al.
1986, Hamilton et al. 1992; Forsberg et al. 1993;
Winemiller and Akin unpublished).
and implications for fisheries management 291
Table 1: Estimated trophic positions of dominant piscivores in floodplain river ecosystems and estuaries (References are 1-
Winemiller 1990, 2- Peterson 1997, 3- Jepsen & Winemiller 2002, 4-Winemiller 1996a, 5- Akin 2001, 6- Winemiller & Akin
unpublished data).
Piscivore Trophic position Site Analysis method Reference
Pygocentrus cariba Valenciennes 3.4 Caño Maraca, Venezuela diet 1
Hoplias malabaricus (Bloch) 3.4 Caño Maraca, Venezuela diet 1
Caquetaia kraussii (Steindachner) 3.5 Caño Maraca, Venezuela diet 1
Cichla orinocensis Humboldt 4.0 Morichal Charcote,Venezuela diet 2
Hoplias malabaricus 4.0 Morichal Charcote,Venezuela diet 2
Cichla orinocensis 3.5 Cinaruco River,Venezuela isotopes 3
Cichla temensis 3.6 Cinaruco River,Venezuela isotopes 3
Cichla temensis 4.8 Pasimoni River, Veneuela isotopes 3
Serrasalmus manueli 3.8 Cinaruco River, Venezuela isotopes 3
Fernandez-Yepez & Ramñrez
Pygocentrus cariba 3.8 Apure River, Venezuela isotopes 3
Hoplias malabaricus 3.6 Apure River, Venezuela isotopes 3
Hoplias malabaricus 4.0 Aguaro River, Venezuela isotopes 3
Hydrolycus armatus (Schomburgk) 3.6 Apure River, Venezuela isotopes 3
Hydrolycus armatus 4.2 Aguaro River, Venezuela isotopes 3
Hydrolycus armatus 3.7 Cinaruco River, Venezuela isotopes 3
Pseudoplatystoma fasciatum (L.) 3.5 Apure River, Venezuela isotopes 3
Pseudoplatystoma fasciatum 4.4 Pasimoni River, Venezuela isotopes 3
Nandopsis dovii (Gñnther) 3.3 Tortuguero River, Costa Rica diet 1
Gobiomorus dormitor (Lacepede) 3.3 Tortuguero River, Costa Rica diet 1
Hepsetus odoe 4.3 Zambezi River, Zambia diet 4
Hydrocynus vittatus 4.6 Zambezi River, Zambia diet 4
Serranochromis robustus (Gñnther) 3.7 Zambezi River, Zambia diet 4
Lepisosteus osseus (L.) 3.6 Brazos River, Texas diet 4
Lepisosteus oculatus (Winchell) 3.3 Brazos River, Texas diet 4
Lepisosteus oculatus 3.3 Mad Island Marsh, Texas diet 5
Lepisosteus oculatus 3.1 Mad Island Marsh, Texas isotopes 6
Sciaenops ocellatus (L.) 3.4 Mad Island Marsh, Texas diet 5
Sciaenops ocellatus 3.3 Mad Island Marsh, Texas isotopes 6
Mean 3.7

Descriptions of food web structure in river-
floodplain ecosystems based on analysis of both diets
and stable isotopes reveal short food chains. In terms
of biomass, the most important pathways connect
detritus to detritivorous fishes (and to a lesser extent
invertebrates) and to piscivorous fishes. Consumer
trophic positions can be estimated as a continuum
using algorithms applied to dietary or isotopic data. In
river-floodplain systems, large abundant piscivores
almost invariably occupy positions between the third
and fourth trophic levels (Table 1). This pattern arises
because piscivore diets are dominated by detritivores
and other fishes feeding near the second trophic level.
In Caño Maraca, a creek-floodplain ecosystem in the
Venezuelan llanos, the most abundant species in the
fish assemblage, Steindachnerina argentea (Gill), also
was the dominant prey of abundant red-belly piranhas
(Pygocentrus cariba) and guavinas (Hoplias malabar-
icus) (Winemiller 1990). In the Cinaruco River, detri-
tivorous and algivorous hemiodid and prochilodontid
fishes dominate the diet of abundant Cichla temensis
(Jepsen et al. 1997). In the Apure River, detritivorous
Prochilodus mariae Eigenmann dominate the diet of
the two most abundant large catfishes,
Pseudoplatystoma fasciatum and P. tigrinum
(Valenciennes) (Barbarino and Winemiller unpub-
lished). Clearly, most matter and energy passes from
the base to the top of the aquatic food web via food
chains that are short (2-3 links and 3-4 levels). Isotopic
analysis of fishes in a Pantanal lake indicated 3-4
trophic levels, with consumers arranged along a troph-
ic continuum rather than discrete levels (Wantzen,
Machado, Voss et al. 2002). Lewis et al. (2001) noted
that short food chains facilitate efficient transfer of
energy from algae to fishes and may explain why large
fish stocks in tropical floodplains can be supported by
the minor algal component of system primary produc-
tion.
Given the dominant role of a relatively small
number of short food chains, the high complexity of
river-floodplain food webs is derived from numerous
weak links among diverse species of both common and
rare taxa. The most numerically abundant species (e.g.
algae, invertebrates, fishes) are small-bodied with low
to moderate standing stocks of biomass. Given high
rates of population turnover, many of these taxa prob-
ably have greater functional significance in food webs
than their low abundance implies. Although average
food chain length leading to top piscivores is short, this
does not imply that all food chains are short. Longer
chains involving small or rare species can be identi-
fied. Small fishes that consume scales, fins, mucus, or
blood of other fishes occur in most large rivers of
South America. These fishes represent insignificant
components of system biomass, but they contribute to
high species diversity and high food web complexity.
Thus, longer food chains that involve small or rare
species are common and increase ecological complex-
ity, but probably have very minor effects on primary
and secondary production. In terms of biomass, tropi-
cal river food webs appear to consist of dominant
(foundation, or core) species connected by short food
chains, plus a much richer assemblage of small (subor-
dinate, or interstitial) species, many of them uncom-
mon, that greatly increase food web complexity while
having relatively little influence on material and ener-
gy flow with the ecosystem. Of course these species
could have important ecological functions that have
not yet been identified (e.g. seed dispersal for riparian
plants, Goulding 1980).
SPECIES FUNCTIONAL DIVERSITY IN
LARGE RIVER FOOD WEBS
The tropics are widely recognized to harbour
higher taxonomic and ecological diversity than tem-
perate regions and large river systems provide no
exception to this rule. Globally, fish species richness is
strongly related to basin size (Welcomme 1985;
Oberdorff, Guegan and Hugueny 1995). However,
fishes show greater taxonomic and ecological diversi-
ty in lowland continental rivers of tropics relative to
comparable rivers of temperate regions (Winemiller
1991a). Whereas the core feeding groups are repre-
sented in both temperate and tropical regions (i.e. algi-
vores, detritivores, omnivores, invertivores and pisci-
vores), the relative proportions differ. Fish assem-
blages of large tropical rivers contain greater fractions
of detritivorous, herbivorous and omnivorous fishes
relative to temperate fish assemblages (Winemiller
292 Floodplain river food webs: Generalizations

1991a). In this regard, tropical river fishes appear to
occupy niche space occupied by invertebrates in tem-
perate rivers.
Although no formal comparisons appear to
have been made, macroinvertebrate species richness in
large rivers does not seem to reveal a latitudinal gradi-
ent as steep as that of fishes. Bivalve mollusks actual-
ly have greater species richness in temperate rivers of
the Western Hemisphere and the abundance and func-
tional diversity of aquatic insects in lowland rivers
does not appear to be much greater in tropical than
temperate rivers. In tropical blackwater rivers (high
concentrations of dissolved organic compounds, low
PH and conductivity, low concentrations of nutrients
and suspended solids), aquatic insect abundance is low
with most species and biomass concentrated in leaf lit-
ter and woody debris. Shrimp are abundant in most
lowland tropical rivers, with various taxa feeding on
detritus, algae and microfauna. Even oligotrophic trop-
ical blackwater rivers can support large populations of
atyid and palaemonid shrimp. Leaf litter and woody
debris seem to provide particularly important habitats
in blackwater rivers (Benke et al. 1984). In tropical
whitewater rivers (high concentrations of nutrients and
suspended sediments in flowing channels, high con-
ductivity, neutral pH), the root zone of floating aquat-
ic macrophytes, such as Paspalum repens and
Eichhornia spp., support high biomass of aquatic
macroinvertebrates. Macroinvertebrates in channel
habitats are concentrated in patchy, structurally com-
plex habitats, such as woody debris (Benke et al.
2001). Clay nodules at the bottom of deep channel
areas of Neotropical whitewater rivers support mayfly
populations that consume detritus and provide a major
food resource for weakly-electric gymnotiform fishes
(Marrero 1987). Gymnotiforms also feed heavily on
planktonic microcrustacea that feed on phytoplankton
(Lundberg et al. 1987).
As noted above, a relatively small fraction of
the total species in a community appear to have large
roles in the flow of matter and energy in floodplain
river food webs. Yet species affect ecosystem proper-
ties via mechanisms besides consumer-resource inter-
actions. Some of the most dominant species of large
lowland rivers have been shown to have strong effects
on ecosystem structure and processes. Afew benthivo-
rous fish species have been shown to disproportionate-
ly influence sediments of channel or floodplain habi-
tats. Using field experiments, Flecker (1996) showed
how benthivorous Prochilodus mariae remove organ-
ic-rich sediments and change the structure of benthic
algae and insect assemblages in a whitewater river of
the Andean piedmont in Venezuela. Semaprochilodus
kneri have similar effects in clearwater and blackwater
rivers in Venezuela (Winemiller unpublished). North
American gizzard shad (Dorossoma cepedianum
(Lesueur) feed on detritus and move nutrients from
sediments to the water column in reservoirs (Vanni
1996). The gizzard shad is a common detritivore and
periphyton grazer of lowland rivers in North America
and could significantly affect ecosystem dynamics.
Benthic feeding by large omnivorous cypriniform fish-
es (e.g. Ictiobus spp., Cyprinus carpio L.) can increase
sediment suspension in the water column (Drenner,
Smith and Threlkeld 1996). Other grazing taxa have
been shown to affect standing stocks of algae and
organic sediments in tropical and temperate rivers.
Field manipulations have shown grazer effects on
standing stocks of algae and organic sediments in
upland tropical and temperate rivers, including studies
involving shrimp (Crowl et al. 2001), tadpoles
(Flecker, Feifarek and Taylor 1999) and aquatic insect
larvae (Power 1990, 1992).
In tropical lowland rivers, a few predatory
species may disproportionately influence the distribu-
tion or abundance of prey populations. Jackson (1961)
proposed that tigerfish (Hydrocynus spp.) restrict use
of main channels of African rivers to a subset of the
fish fauna that possess morphological features that
inhibit predation (e.g. deep body, dorsal and pectoral
spines). In South American rivers, piranhas appear to
restrict the use of open-water off-shore areas by many
fishes (Winemiller 1989a). Experimental exclusion of
Cichla species and other large piscivores significantly
affected the abundance and size distribution of fishes
in the Cinaruco River, Venezuela (Layman and
Winemiller unpublished).
and implications for fisheries management 293

FOOD WEB DYNAMICS IN RESPONSE TO
FLOOD PULSES
E
FFECT OF THE FLOOD PULSE ON PRODUCTION DYNAMICS
The temporal dynamics of disturbance and
recovery of local habitats in the river-floodplain habi-
tat mosaic drive spatiotemporal variation in primary
production sources and favour alternative life history
strategies. According to the flood-pulse model, flood
conditions should be associated with greater nutrient
availability, aquatic primary production (dominated by
macrophytes), allochthonous inputs and secondary
production, especially among juvenile fishes, in flood-
plain habitats. Low-water conditions result in contrac-
tion of marginal aquatic habitats, death and decay of
aquatic macrophytes and higher densities of aquatic
organisms, including phytoplankton and zooplankton
in floodplain lagoons (Rai and Hill 1984; Putz and
Junk 1997). Because overall productivity is lower dur-
ing low-water conditions and densities of consumer
taxa are high, there is a strong advantage for spawning
during flood pulses, but only if these pulses endure
long enough to yield sufficient survival and growth of
early life stages prior to flood subsidence.
In a strongly seasonal environmental regime,
species have the opportunity to evolve adaptations to
exploit relatively predictable habitats and resources
(Southwood 1977, Winemiller and Rose 1992, Resh et
al. 1994). Under this regime, a periodic life history
strategy is favoured (i.e. seasonal spawning, high
fecundity, small eggs and larvae, little parental care).
In tropical-seasonal systems, temperature is relatively
constant and periodic flooding is the primary factor
driving ecological dynamics. Access to floodplain
habitats is important for successful recruitment by
many fish species in tropical-seasonal rivers. Inter-
annual variation in fish recruitment generally is more
strongly associated with flood duration than flood
magnitude. In the Upper Paraná floodplain-river sys-
tem, years with higher and longer duration floods were
associated with increases in condition, growth and
recruitment of Prochilodus scrofa Steindachner
(Gomes and Agostinho 1997). In tropical northern
Australia, fish abundance in billabongs (oxbows) was
positively correlated with duration of the annual flood
(Madsen and Shine 2000). Even so, a range of success-
ful life-history strategies is observed among fish
species of tropical lowland rivers (Winemiller 1989b,
1996a, 1996b). Small opportunistic species with high
reproductive effort protracted spawning periods and
short-life spans are common in shallow marginal habi-
tats that are constantly shifting across the river-flood-
plain landscape as water level rises and falls. The most
extreme examples of the opportunistic strategy are
observed among annual killifishes (Aplocheilidae) that
inhabit shallow ephemeral pools. Many equilibrium
strategists (relatively low fecundity with well-devel-
oped parental care) spawn just prior to the annual flood
pulse and then move into newly flooded areas to
brood. Based on growth variation, this seasonal
spawning pattern seems to apply to Cichla species in
Venezuela (Jepsen et al. 1999) and Serranochromis
species in the Upper Zambezi River (Winemiller
1991b). Fishes with the equilibrium strategy may have
higher reproductive success when water fluctuation is
low. Some of the brood-guarding species of the upper
Paraná River have greater abundance during years with
low floods (Agostinho et al. 2000).
In temperate-seasonal rivers, access to flooded
habitats may be non-essential, beneficial but non-
essential, or detrimental to recruitment. Flooding
enhances nutrient concentrations; particle loads and
phytoplankton biomass in connected floodplain habi-
tats (Hein et al. 1999), but can reduce densities of crus-
tacean zooplankton (Baranyi et al. 2002). In temperate
regions, temperature may have an influence on repro-
ductive strategies that is equal to or greater than flood-
ing. When warming temperatures coincide with a reli-
able annual flood pulse, selection should favour a peri-
odic strategy just as in the tropics. Indeed, contracted
spawning of large batches of small eggs is the domi-
nant pattern observed in temperate-seasonal river fish
faunas. Greater availability of floodplain habitats
enhances fish recruitment and species diversity in low-
land rivers in Europe (Copp 1989; Schiemer et al.
2001a) and North America (Sparks 1995). As in tropi-
cal systems, other life history strategies succeed in
temperate-seasonal systems (e.g. sunfishes with rela-
294 Floodplain river food webs: Generalizations

tive equilibrium strategies and small cyprinids and
poeciliids with opportunistic strategies). Humphries,
King, and Koehn (1999); Humphries et al. 2002) iden-
tified three fish life-history strategies (gradient similar
to model of Winemiller and Rose 1992) among fishes
of Australia’s Murray-Darling system. Flood regimes
of many rivers of this region are regulated.
Unregulated rivers display a temperate-seasonal pat-
tern (Figure 1) but with large inter-annual variation in
the magnitude of the seasonal flood-pulse. Humphries
and co-workers discovered that virtually all fish
species spawn each year with variable recruitment suc-
cess depending on flow and temperature conditions.
Because large floods do not occur each year, many
species are able to recruit successfully by spawning
and completing their life cycle entirely within main-
channel habitats (the “low flow recruitment hypothe-
sis”). Their studies demonstrate the potential impor-
tance of marginal channel habitats with low current
velocity and abundant benthic micro-invertebrates that
support fish early life stages.
In aseasonal flood-pulse regimes, aquatic
organisms are more challenged to respond appropriate-
ly to relatively unpredictable patterns of resource vari-
ation. As in the Murray-Darling system, spatiotempo-
ral connectivity of habitats and access to floodplain
habitats is nonessential for most species, but greatly
enhances recruitment for many, if not most, species in
temperate-aseasonal rivers. Winemiller et al. (2000)
discovered that certain fish species dominated oxbow
lakes and others were more common in the active
channel of the Brazos River, Texas. Opportunistic
species numerically dominated the river channel and
shallow oxbow lakes with high rates of disturbance
and periodic strategists dominated deeper oxbow lakes
with irregular but periodic flood connections to the
river (Winemiller 1996a). When flooding occurs dur-
ing springtime, recruitment by periodic strategists,
such as gizzard shad, buffalo (Ictiobus bubalus
(Rafinesque)) and crappie (Pomoxis annularis
Rafinesque) is high. Yet springtime floods only occur
during some years, so that spawning during most years
is associated with low recruitment success (Winemiller
unpublished data). Interspecific differences in respons-
es to hydrologic regimes in habitats across the lateral
floodplain gradient have been shown for other taxo-
nomic groups in other regions, including trees (Junk
1989), phytoplankton (van den Brink et al. 1993) and
benthic macroinvertebrates (Marchese and Ezcurra de
Drago 1992).
EFFECT OF THE FLOOD PULSE ON CONSUMPTION
DYNAMICS
The expansion and contraction of aquatic habi-
tats in response to flooding has a major influence on
consumer-resource interactions. Newly expanded
floodplain habitats provide an immediate influx of
allochthonous detritus and invertebrates and, with
time, greater nutrient availability and aquatic primary
production. Densities of aquatic organisms are low ini-
tially and increase over time as new individuals recruit
under productive flood conditions. Fish growth rate
and condition are high in flooded habitats (Welcomme
1985). In the central Amazon, juveniles of omnivorous
species, but not detritivorous species, grew faster dur-
ing the rising-water period (Bayley 1988). Growth of
omnivores was positively associated with flood magni-
tude and in all cases growth appeared to be density-
independent.
Highest fish abundance and per-unit-area den-
sities typically occur as floodwaters recede. As dictat-
ed by the functional response, the falling-water period
is when predator-prey interactions are most intense.
This is also the period when resource limitation may
occur for species that exploit algae and aquatic and ter-
restrial invertebrates. Bayley (1988) found that juve-
niles of only 2 of 8 omnivorous species in the central
Amazon showed significant evidence of density-
dependent growth during the falling water period. For
piscivores, the falling-water period represents a time of
resource abundance, as fishes become increasingly
concentrated in aquatic habitats of reduced volume.
Piscivore feeding rates increase during the falling
water period and piscivore growth and body condition
increase (Jepsen et al. 1999). If piscivores deplete prey
populations during the falling-water period, they may
eventually become resource limited for several months
during the lowest water stages. For size-selective
and implications for fisheries management 295

(gape-limited) piscivores, optimal prey sizes become
depleted first and piscivores shift to increasingly
smaller prey as water levels continue to fall. Jepsen et
al. (1997) described a decline in mean prey size con-
sumed by Cichla species in the Cinaruco River during
the 6-month falling water period. This shift in the aver-
age size of consumed prey size almost exactly match-
es the shift in the mode for the size distribution of fish-
es in the littoral zone (Layman and Winemiller unpub-
lished data).
The scope of seasonal changes in population
densities and predator-prey interactions obviously
depends on the timing, magnitude and duration of
flooding. The scope of these changes will be smaller in
temperate-aseasonal rivers and greater in seasonal
rivers with floras and faunas well adapted to take
advantage of periodic changes in habitat and resource
quality and availability. As a result, seasonal rivers can
sustain greater fish harvest than aseasonal rivers in
landscapes with comparable geomorphology and nutri-
ent availability. Power et al. (1995) created a simple
simulation model that linked floodplain river hydrolo-
gy to food web dynamics based on the Lotka-Volterra
algorithms. They examined four scenarios: a river with
connection to its floodplain and seasonal (sinusoidal)
discharge, a river confined by levees with sinusoidal
discharge and regulated rivers with low and average
discharge that never lead to flooding. Only the con-
nected river with seasonal discharge produced stable
populations of predators and grazers. The leveed river
yielded unstable predator-prey dynamics as a result of
channel confinement and regulated rivers resulted in
low or oscillating grazer populations that ultimately
were unable to sustain viable predator populations.
Whereas this model represents a gross oversimplifica-
tion of natural food webs, the findings highlight the
influence of discharge dynamics and channel-flood-
plain connections on community dynamics.
EFFECTS OF THE FLOOD PULSE ON MIGRATION
In addition to its effects on population dynam-
ics and consumer-resource interactions, flooding also
influences movement of materials and organisms.
Movement in response to flooding may be essentially
longitudinal or lateral and passive or active. Seasonal
succession and food web dynamics are influenced by
all of these forms of movement. The initial stages of a
flood pulse submerge terrain which results in inputs of
dissolved inorganic nutrients from terrestrial vegeta-
tion, both living and dead (Junk et al. 1989). Surface
runoff and floodwater recession carries these nutrients
into channel areas where aquatic production may be
stimulated (Rai and Hill 1984; Putz and Junk 1997;
Lewis et al. 2000). Likewise, phytoplankton, zoo-
plankton, floating macrophytes and terrestrial
allochthonous resources are washed into flowing chan-
nels as well as deeper permanent floodplain lagoons.
Based on a mass-balance approach, Lewis et al. (2000)
concluded that the floodplain of the lower Orinoco
River exports no organic carbon to the river channel.
They concluded that this hydrologically open system
behaves like a closed system with respect to organic
carbon balance. They observed that the natural levee of
the floodplain restricts water movement to a direction
parallel to the longitudinal axis of the river channel.
Thus, passive export of organic carbon is low because
only a minor fraction of water actually passes from the
floodplain to the channel. Presumably then, flood-
plains internally recycle organic carbon captured from
surrounding uplands.
The Lewis et al. (2000) carbon-balance model
does not consider active movement by aquatic organ-
isms. Fishes, in particular, migrate between channel
and floodplain locations in response to seasonal
changes in the relative benefits and costs associated
with conditions in each area (Welcomme 1985).
Flooding provides fishes with almost unlimited access
to a range of habitats. In tropical-seasonal rivers, fish
movements from river channels into floodplain habi-
tats are particularly regular (Goulding 1980;
Welcomme 1985; Fernandes 1997; Hocutt and
Johnson 2001). In temperate-seasonal and temperate-
aseasonal rivers, these fish movements are common,
but apparently less predictable. Depending on the
taxon and region, tropical river fishes may migrate
locally (1-100 km) or regionally (>100 km). In the
llanos region of the Orinoco Basin, many and probably
most, fishes perform local migrations into seasonally
296 Floodplain river food webs: Generalizations

inundated savannas for reproduction. These seasonal
habitats are highly productive and serve as classic
nursery areas that enhance juvenile growth and sur-
vival (Winemiller 1989b, 1996b). When water levels
drop, these areas become hypoxic and fishes that fail to
migrate downstream to deeper channels risk death
from hypoxia or stranding in drying pools (Lowe-
McConnell 1964). Even though many floodplain fish-
es possess special adaptations for dealing with aquatic
hypoxia (Kramer et al. 1978), a great deal of aquatic
biomass moves out of floodplain habitats into deeper
creeks and rivers. During the annual falling-water peri-
od, piscivores in mainstem rivers feed heavily on fish-
es that migrate out of tributaries draining the flood-
plains (Winemiller 1996a; Winemiller and Jepsen
1998). Thus, if we add these higher food web compo-
nents to Lewis et al. (2000) calculation of organic car-
bon mass-balance, floodplains export large amounts of
organic carbon to river channels.
Some river fishes undergo regular seasonal
migrations on regional scales. Welcomme (1985) sum-
marized evidence of longitudinal and lateral migra-
tions by South American and African fishes. Highly
migratory fishes can be extremely abundant with
strong effects on local food webs. In rivers of the North
Pacific region, the decaying carcasses of anadromous
salmon import significant amounts of limiting nutri-
ents that can enhance ecosystem productivity during
summer (Kline et al. 1990; Willson, Gende and
Marston 1998; Cederholm et al. 1999). In South
American rivers, prochilodontid and other characiform
fishes perform seasonal migrations of hundreds of
kilometres (Bayley 1973; Vazzoler, Amadio and
Daraciolo-Malta 1989; Ribeiro and Petrere 1990).
Immigration of these abundant fishes during the
falling-water period produces large effects on local
food webs. First, prochilodontids have large effects on
sediments and ecological dynamics in benthic commu-
nities (discussed above). Thus, prochilodontids are
both ecosystem engineers as well as strong interactors
with benthic elements of the food web (Flecker 1996).
Second, immigrating prochilodontids provide an abun-
dant food resource for resident piscivores (discussed
above), which can be particularly significant for olig-
otrophic systems that receive young migrants from
more productive systems. In this capacity,
prochilodontids provide a spatial food web subsidy
(Polis, Anderson and Holt 1997), in which material
from a more productive ecosystem (floodplain wet-
lands) enters the food web in a less productive ecosys-
tem (flowing channel). Food web subsidies can have
major effects on food web dynamics, including induc-
tion of trophic cascades (Polis et al. 1997; Winemiller
and Jepsen 2002) and stabilization of complex systems
(Huxel and McCann 1998).
Some large predatory fishes of floodplain
rivers also undergo long-distance regional migrations.
Barthem and Goulding (1997) described migrations by
large pimelodid catfishes that span almost the entire
Amazon Basin. African tigerfish (Hydrocynus spp.),
Alestes and Labeo species migrate longitudinally
according to seasonal hydrological regime (Jackson
1961; Welcome 1985). Predatory ariid, centropomid
and eleotrid fishes of Australia, Southeast Asia, the
East and West Indies and tropical Americas habitually
migrate between rivers and coastal marine waters. The
food web implications of these “reverse subsidies”
have scarcely been explored. If the effects of exotic
piscivores on lake communities (Zaret and Paine 1973;
Kaufman 1992) provide any indication, the effects of
immigrant piscivores on fish populations in local flu-
vial habitats are potentially great. Likewise, removal
of resident piscivores can affect local populations.
Negative impacts of commercial fishing on large pisci-
vores in floodplain lagoons of the Cinaruco River had
a significant effect on local assemblage structure of
small prey fishes (Layman and Winemiller unpub-
lished).
MANAGEMENT IMPLICATIONS OF FOOD
WEB ECOLOGY
Floodplains of lowland rivers provide impor-
tant ecosystem services (i.e. nutrient cycling, flood
mitigation) and renewable natural resources (e.g. fish-
ery and forest products). Human impacts on river-
floodplain systems have been described repeatedly
(Welcomme 1985; Ward and Stanford 1989; Bayley
1995; Sparks 1995; Dudgeon 2000; Pringle, Freeman
and implications for fisheries management 297

and Freeman 2000), but the focus of discussion here
will be the interaction between food web ecology,
human impacts and sustainable fisheries.
HABITAT CONNECTIVITY
Dams obviously fragment rivers in the longitu-
dinal dimension. Many important river fishes undergo
seasonal longitudinal migrations that make them high-
ly vulnerable to impacts from not only dams, but also
other channel obstructions such as weirs and gillnets.
As discussed above, some of these fishes have large
ecosystem effects (e.g. salmon affecting nutrients). In
addition to affecting sediments and benthic biota,
migratory prochilodontids also provide nutritional sub-
sidies to piscivores that likely affect food web dynam-
ics in the receiving communities.
A major human impact on large rivers is levee
construction for the purpose of preventing floodplain
inundation or draining of wetlands for agriculture and
other land uses. Levees obviously disrupt important
connections between river channels and floodplains,
which cuts off exchanges of material and organisms
among dynamic habitats critical for completion of
species life cycles (Ward et al. 1999; Amoros and
Bornette 2002) and ecosystem dynamics (Junk et al.
1989; Aspetsberger et al. 2002). Disconnecting the
river channel from its floodplain has obvious negative
impacts on nutrient cycling (Tockner et al. 1999), sys-
tem productivity (Bayley 1989; Junk et al. 1989;
Agostinho and Zalewski 1994) and biodiversity
(Schiemer et al. 2001a; Robinson et al. 2002).
Magnitudes of these impacts should be greater for
tropical- and temperate-seasonal rivers than for tem-
perate-aseasonal rivers. For example, recruitment by
fishes in temperate-aseasonal rivers usually is more
dependent on temperature regime than flood regime.
Reproductive timing and recruitment by fishes in trop-
ical floodplain rivers are strongly correlated with
dynamics of the annual flood pulse. Large cichlids in
South America (Cichla, Hoplarchus, Heros spp.) and
Africa (Serranochromis, Oreochromis spp.) exhibit
protracted spawning periods in reservoirs, but season-
al, contracted spawning periods in rivers (Winemiller
personal observation). Evidence from temperate rivers
indicates that many fish species complete their entire
life cycle within the main channel (Galat and
Zweimüller 2001; Dettmers et al. 2001) although even
these species are strongly dependent on natural flood
regimes (Schiemer et al. 2001b). Early life stages of
these lotic-adapted species frequently depend on
nearshore channel habitats with relatively lentic condi-
tions. The inshore retention of fish larvae and their
food resources is a critical feature influenced by river
geomorphology and hydrology (Schiemer et al.
2001b).
Human impacts that reduce habitat connections
in river-floodplain landscapes also can affect biodiver-
sity and food webs by inhibiting patch colonization
and community succession (Sedell et al. 1990). Recent
research on the Cinaruco River in Venezuela indicates
that fishes and macroinvertebrate communities of the
littoral zone are significantly structured in relation to
substrate type (Arrington and Winemiller unpub-
lished). Habitat patches are colonized and abandoned
in sequence as they are submerged and exposed by the
moving littoral zone. Field experiments demonstrated
that artificial habitat patches undergo community suc-
cession that is accompanied by increasing degrees of
non-random assemblage structure (Winemiller et al.
unpublished). The littoral food web appears to con-
form to Holt’s (1996) spatial model of food web
dynamics. In this model, taxa at lower trophic levels
are restricted to the smallest habitat patches, with larg-
er, more mobile consumers at higher trophic levels
feeding across multiple patches. This pattern continues
in a trophic hierarchy that ultimately yields a sink web
defined by food chains terminating with a single large,
mobile top predator. River channelization, levee con-
struction and wetland drainage disrupt not only com-
munity dynamics in the littoral zone, but also restrict
access by predators to habitat patches containing prey
(Toth et al. 1998). Disruption of both factors (commu-
nity assembly and predation by large mobile fishes) is
certain to affect biodiversity.
Fishes are not the only vertebrates that depend
on dynamic connections between channel and flood-
plain aquatic habitats. Dynamic habitats of river-flood-
298 Floodplain river food webs: Generalizations

plain systems enhance species diversity of aquatic
insects (Smock 1994), mussels (Tucker, Theiling and
Camerer 1996), turtles (Bodie and Semlitsch 2000),
birds (Remsen and Parker 1983) and mammals
(Sheppe and Osborne 1971).
FLOW REGIMES
Regulation of river hydrology changes natural
flood regimes that determine elemental cycles, system
productivity, reproduction and population dynamics of
aquatic organisms and consumer-resource interactions.
Clearly, significant alteration of the natural flood-
regime in temperate- and tropical-seasonal rivers will
have detrimental effects for native fish species that
time reproduction to maximize recruitment success
under predictable patterns of spatio-temporal environ-
mental variation. High primary production and inputs
of allochthonous resources that accompany flood-puls-
es tend to enhance fish recruitment success, but some
species are less responsive than others. Many species
achieve low to moderate recruitment even under no-
flow conditions (Humphries et al. 2002).
Consequently, community dynamics are partially a
function of the timing and magnitude of flooding and
this is bound to have large effects on food web dynam-
ics that in turn influence dynamics of exploitable fish
stocks. For example, years in which the Upper Paraná
River, Brazil experiences higher, longer duration
floods produce greater abundance of age-0
Prochilodus scrofa, the most important commercial
fish of the region (Gomes and Agostinho 1997).
Prochilodus is a principal prey for Salminus maxillo-
sus Valenciennes, Plagioscion squamosissimus
(Heckel) and other large piscivores that are important
in the local fishery (Hahn et al. 1997). Thus, flood
pulses affect these large predators both directly, in
terms of their own recruitment success, as well as indi-
rectly via food chain interactions. Management of mul-
tispecies fisheries in large rivers requires a food web
perspective. Stock dynamics are influenced both by
bottom-up factors related to ecosystem productivity
and by top-down factors influenced by relative densi-
ties of predator and prey populations.
and implications for fisheries management 299
Flood dynamics affect both bottom-up and top-
down effects in food webs. In large tropical rivers,
flooding occurs predictably over large areas, which
results in a pulse of primary production (Junk et al.
1989). This, in turn, is efficiently transferred to higher
trophic levels due to species life history strategies that
maximize fitness (i.e. population rate of increase)
under predictable regimes of environmental variation.
Harvest rates increase as fish populations become vul-
nerable to fishing when flood subsidence increases
their per-unit-area densities (i.e. a functional
response). The world’s most productive river fisheries
are associated with seasonal flood-pulse dynamics in
tropical areas. Holding all other factors equal, nutrient-
rich landscapes in the tropics (e.g. Mekong, Niger,
Zambezi, middle Orinoco and lower Amazon rivers)
produce greater fish yields than nutrient-poor regions
(Rio Negro and other rivers draining South America’s
Guyana Shield region). In temperate regions, lower
temperatures result in lower annual productivity. On
geologic-evolutionary time scales, temperate regions
have experienced more recent and frequent climatic
disturbances that have inhibited biological diversifica-
tion and ecological specialization within regional fish
faunas. Currently, there is much interest in the poten-
tial positive relationship between biodiversity and
community productivity (e.g. Tilman 1999) and this
relationship could contribute to the greater productivi-
ty of seasonal tropical-seasonal river fish assemblages
relative to those of temperate-seasonal rivers.
Fish production should be lowest in temperate-
aseasonal rivers for three reasons. The timing of floods
often will not coincide with periods with highest tem-
peratures. Additionally, the timing of floods often will
not synchronize with the spawning periods innately
cued to photoperiodicity and seasonal temperature
variation. Finally, temperate faunas are less likely to
have evolved life history strategies and ecological
adaptations designed to capitalize on flood pulse con-
ditions, because these conditions are unpredictable on
both intra- and inter-annual time scales. All other fac-
tors being equal, temperate-aseasonal rivers are less
resistant to intense sustained harvest, of the kind prac-
ticed for generations in many tropical regions.

Direct consumption of allochthonous resources
by fishes is particularly important in forested lowland
regions of the Amazon Basin, with some species
notably adapted for consuming fruits and seeds
(Goulding 1980; Loubens and Panfili 2001). Reduced
flood frequency, in addition to deforestation, will neg-
atively impact direct entry of allochthonous resources
into aquatic food webs, to the detriment of yields of
several commercially important stocks (Goulding
1980; Reinert and Winter 2001).
On geological time scales, flood regimes main-
tain physical habitat heterogeneity by alternately erod-
ing and depositing sediments on the landscape
(Kellerhals and Church 1989). On shorter time scales,
erosion and deposition of sediments are disturbances
to vegetation communities. Natural hydrological
processes create new substrates for community succes-
sion. The result is a rich mosaic of habitat patches with
different degrees of structural complexity, exposure to
natural disturbances and community composition
(Shiel, Green and Neilsen 1998). Thus, chronic
absence of flooding results in altered disturbance
regimes and ultimately lowers habitat heterogeneity
and species diversity (Schiemer at al. 2001a).
Flow regimes, in concert with soils and land-
scape geomorphology, also influence suspended sedi-
ment loads. Turbidity influences predatory-prey inter-
actions and community composition and dynamics.
Highly turbid systems often are dominated by siluri-
form fishes and, in Africa and South America respec-
tively, weakly electric fishes (mormyriforms and gym-
notiforms). Predators that rely on vision, such as cich-
lids and many characiform and cypriniform fishes,
tend to be scarce in turbid whitewater rivers. In turbid
river-floodplain systems, visually orienting fishes are
most abundant in clear tributaries creeks and lacustrine
habitats of floodplains where sediments settle out.
Turbidity varies among floodplain lagoons as a func-
tion of local soils and other landscape features. During
the dry season, water transparency is associated with a
fairly consistent pattern of fish assemblage composi-
tion in Orinoco River floodplain lagoons, with turbid
lagoons having more siluriforms and gymnotiforms
and clear lagoons having more characids (Rodríguez
and Lewis 1997). Wet-season flooding mixes water
and allows organisms to move freely across the land-
scape, which presumably homogenizes these lagoon
fish assemblages. The effect of turbidity on river food
web structure and dynamics has not been investigated.
FISHERIES HARVEST
Fisheries obviously impact river food webs in
many different ways. Overfishing changes consumer-
resource dynamics and the distribution of interaction
strengths in the food web. If affected populations are
species with large functional importance to the com-
munity or ecosystem, the effect of their depletion may
be large and immediate. For example, overharvest of
benthivorous prochilodontids would fundamentally
alter the sediment dynamics and benthic ecology in
Andean piedmont rivers. There is some evidence that
this is already occurring in Venezuela where extensive
gillnetting removes large numbers of Prochilodus
mariae during their upstream migrations (Barbarino-
Duque, Taphorn and Winemiller 1998). With reduced
densities of Prochilodus that consume and resuspend
fine sediments, river channels accumulate a thick layer
of soft sediments that inhibit development of a benthic
community dominated by diatoms and grazing insects
(Flecker 1996). Because benthic primary production is
the principal energy source in this system, the entire
food web undoubtedly changes with unknown conse-
quences for biodiversity and secondary production.
Similar effects of prochilodontids on benthic process-
es have been demonstrated experimentally in channel
and lagoon habitats of the Cinaruco River (Winemiller
et al. unpublished data).
In North America and Europe, commercial
fishing in rivers is relatively insignificant. In cold-
water regions, salmonids, esocids and percids are
heavily targeted by sportfishers, sometimes with nega-
tive impacts on stocks. Tropical river fisheries provide
a major source of animal protein for people of devel-
oping countries. Fishing effort in African and Asian
rivers is generally more intense than in South
American rivers, the latter having fisheries that contin-
ue to be dominated by a relatively small number of
300 Floodplain river food webs: Generalizations

large and economically valuable species (Welcomme
1990). Yet some regions of South America have
extremely high fishing effort (Welcomme 1990) and
effort is generally increasing everywhere, in some
cases rapidly. Size overfishing is pervasive in large
rivers worldwide (e.g. Mekong River fisheries dis-
cussed during LARS 2). In Venezuela, maximum and
average sizes of Cichla temensis has declined marked-
ly in rivers over the past 20 years and C. temensis
abundance declined precipitously in the Rio Aguaro
with commencement of commercial netting in the
1970s. The migratory characid Salminus hillari
Valenciennes was a popular sportfish in rivers of the
Andean piedmont of Venezuela until the early 1960s.
The species is now extremely rare due to dam con-
struction and gillnetting (Winemiller, Marrero and
Taphorn 1996). Salminus was once the principal pred-
ator of Prochilodus mariae that migrated en mass into
piedmont rivers during the dry season. Although
Prochilodus also have declined in piedmont rivers
(Barbarino-Duque et al. 1998), this species, unlike
Salminus, has a broad dry season distribution with
large populations maintained in lowland rivers.
Large piscivores often are among the first fish-
es to be targeted by river fisheries. The phenomenon of
“fishing down food webs” was described for marine
systems globally (Pauly et al. 1998). This pattern may
apply equally to river fisheries. In the Amazon, the
abundance and size of pirarucu (Arapaima gigas
(Cuvier) and pimelodid catfishes has declined steadily
in most regions. Although less well documented, a
similar pattern is observed for pimelodid catfishes and
payaras (Hydrolycus spp.) of the Orinoco, Salminus
maxillosus of the Paraná and Lates niloticus (L.) and
Hydrocynus spp. of the Niger, Oeme and other West
African rivers. As stocks of these large piscivores
become depleted, fish markets become even more
strongly dominated by less valuable but more numer-
ous detritivorous and omnivorous species, such as
prochilodontids, Mylossoma and Brycon species in
South America and tilapiine cichlids and Barbus spcies
in Africa. Some of the major predatory fishes inhabit-
ing large warmwater rivers of North America are noc-
turnal catfishes (siluriforms) and lepisosteid gars, the
latter having no commercial value and generating little
sportfishing interest. Because commercial river fish-
eries are insignificant in North America and Europe
and sportfisheries essentially target predatory species,
the fishing-down-food-webs phenomenon has not been
observed in rivers of these regions.
Overharvest of fish stocks changes population
abundance and the structure and dynamics of river
food webs. The elimination of top predators could
yield top-down effects in food chains, but in many
cases prey populations are targeted just as intensely.
Virtually no information is available from any large
river to enable even modest predictions regarding fish-
ing effects on food web dynamics. In tropical rivers,
fish communities are species rich and food webs are
complex. Even when top predators feed on a similar
broad array of prey taxa, fisheries that exploit multiple
predator species can yield chaotic dynamics of individ-
ual populations (Wilson et al. 1991). Fisheries harvest
also can change population size structure, which in
turn affects population dynamics via effects on life his-
tory strategies (e.g. reduction in size at maturity) and
size-dependent predator-prey interactions. These
effects have been demonstrated in fish populations
from streams, lakes and marine systems, but so far lit-
tle information has been gathered from large rivers.
Strong sustained harvest of the largest individuals
selects for earlier age and smaller size of maturation
(Conover and Munch 2002). The combined effects of
overharvest of the largest size classes and the evolution
of smaller size at maturation should profoundly influ-
ence both predator and prey populations when preda-
tion is size-limited. Smaller predators will result in
smaller average and maximum size of consumed prey.
If large piscivores are targeted more intensely than
their prey, as is frequently the case, this could lead to a
negative feedback that affects predator populations
negatively, with potential positive effects on prey
abundance. The study of predator-prey dynamics in
large-river food webs remains in its infancy and a great
deal of research is needed before we can even begin to
construct predictive models.
and implications for fisheries management 301

CONCLUSION
The study of food web ecology in river-flood-
plain systems remains in its infancy. This review has
highlighted only a few of the most basic issues, most
of which are largely unresolved. For example, the
influence of flood regimes on population dynamics of
aquatic organisms with different life history strategies
and regional/evolutionary histories is highly variable.
Therefore, it may be erroneous to assume that regular
flood pulses, of the sort that occur in large tropical
rivers, are required for maintenance of high biodiversi-
ty in every instance. The flood pulse concept of Junk et
al. (1989) probably overestimates the role of flood-
plains for river biota in systems with flood regimes that
are naturally unpredictable or out of phase with spring-
summer. Certainly at some scale of spatial and tempo-
ral resolution, flood pulses are essential for biodiversi-
ty in any river ecosystem. The challenge is to identify
the biological responses to variation at multiple scales.
Food webs are complex and influenced by many abiot-
ic and biotic factors. Although several of the most
important and obvious factors were discussed here,
many more must be examined. For example, exotic
species sometimes dominate river communities (e.g.
European carp in rivers of North America and
Australia), usually with undetermined effects on food
web dynamics and ecosystem processes. Given the
important ecosystem services provided by floodplain
rivers, the high value of river fisheries, especially in
the tropics, as well as the multiple human impacts on
river-floodplain systems, vastly greater research
investment is warranted.
ACKNOWLEDGEMENTS
Rosemary Lowe-McConnell, Craig Layman
and Alexandre Miranda Garcia read earlier drafts of
the manuscript and provided helpful comments. New
information and ideas developed here were aided by
U.S. National Science Foundation Grant 0089834.
302 Floodplain river food webs: Generalizations

REFERENCES
Agostinho A.A. & Zalewski 1994. The dependence of
fish community structure and dynamics on
floodplain and riparian ecotone zone in Paraná
River, Brazil. Hydrobiologia, 303: 141-148.
Agostinho A.A., Thomaz S.M., Minte-Vera, C.V. &
Winemiller K.O. 2000. Biodiversity in the high
Paraná River floodplain. In B. Gopal, W.J.
Junk & J.A. Davis eds. Wetlands Biodiversity.
Leiden, The Netherlands, Backhuys
Publishers. pp. 89-118.
Akin S. 2001. Ecological dynamics of the aquatic com-
munity in a Texas coastal salt marsh. College
Station, TX, USA, A&M University. (Doctoral
dissertation)
Amoros C. & Bornette G. 2002. Connectivity and bio-
complexity in waterbodies of riverine flood-
plains. Freshwater Biology, 47: 761-776.
Araujo-Lima C.A.R.M., Forsberg B.R., Victoria R. &
Martinelli L.A. 1986. Energy sources for detri-
tivorous fishes in the Amazon. Science, 234:
1256-1258.
Arrington D.A. &. Winemiller K. O. 1993. Organization
and maintenance of fish diversity in shallow
waters of tropical floodplain rivers. In R.L.
Welcomme & T. Petr, eds. Proceedings of the
Second International Symposium on the
Management of Large Rivers for Fisheries
Volume 2. Food and Agriculture Organization
of the United Nations & Mekong River
Commission. FAO Regional Office for Asia
and the Pacific, Bangkok. RAP Publication
2004/17. pp. 25-36.
Aspetsberger F., Huber F., Kargl S., Scharinger B.,
Peduzzi P. & Hein T. 2002. Particulate organic
matter dynamics in a river floodplain system:
Impact of hydrological connectivity. Archives
fur Hydrobiologie, 156/1: 23-42.
Baranyi C., Hein T., Holarek C., Keckeis S. & Schiemer
F. 2002. Zooplankton biomass and community
structure in a Danube River floodplain system:
Effects of hydrology. Freshwater Biology, 47:
473-482.
Barbarino-Duque A., Taphorn D.C. & Winemiller K.O.
1998. Ecology of the coporo, Prochilodus
mariae Characiformes, Prochilodontidae, and
status of annual migrations in western
Venezuela. Environmental Biology of Fishes,
53: 33-46.
Barthem R. & Goulding M. 1997. The catfish connection:
ecology, migration, and conservation of ama-
zon predators. New York, Columbia
University Press.
Bayle P.B. 1973. Studies on the migratory characin
Prochilodus platensis Holmberg 1889 Pisces,
Characoidei in the Rio Pilcomayo, S. America.
Journal of Fish Biology, 5: 25-40.
Bayley P.B. 1988. Factors affecting growth rates of young
tropical fishes: Seasonality and density-
dependence. Environmental Biology of Fishes,
21: 127-142.
Bayley P.B. 1989. Aquatic environments in the Amazon
Basin, with an analysis of carbon sources, fish
production, and yield. In D.P. Dodge ed.
Proceedings of the international large rivers
symposium. Can. J. Fish. Aquat. Sci. Spec.
Publ., 106: 399-408.
Bayley P.B. 1995. Understanding large river-flooplain
ecosystems. BioScience, 45: 153-158.
Benedito-Cecilio E., Araujo-Lima C.A.R.M., Forsberg
B.R., Bittencourt M.M. & Martinelli L.C.
2000. Carbon sources of Amazonian fisheries.
Fisheries Management and Ecology, 7: 303-
315.
and implications for fisheries management 303

Benke A.C., VanArsdall T.C., Jr., Gillespie D.M. &
Parrish F.K. 1984. Invertebrate productivity in
a subtropical blackwater river: The importance
of habitat and life history. Ecological
Monographs, 54: 25-63.
Benke A.C., Chaubey I., Ward G.M. & Dunn E.L. 2000.
Flood pulse dynamics of an unregulated river
floodplain in the southeastern U.S. coastal
plain. Ecology, 81: 2730-2741.
Benke A.C., Wallace J.B., Harrison J.W. & Koebel J.W.
2001. Food web quantification using second-
ary production analysis: Predaceous inverte-
brates of the snag habitat in a subtropical river.
Freshwater Biology, 46: 329-346.
Bodie J.R. & Semlitsch R.D. 2000. Spatial and temporal
use of floodplain habitats by lentic and lotic
species of aquatic turtles. Oecologia, 122: 138-
146.
Bunn S.E., Davies P.M. & Winning M. 2003. Sources of
organic carbon supporting the food web of an
arid zone floodplain river. Freshwater Biology,
48: 619-635.
Cederholm C.J., Kunze M.D., Murota T. & Sibatani A.
1999. Pacific salmon carcasses: Essential con-
tributions of nutrients and energy for aquatic
and terrestrial ecosystems. Fisheries, 24: 6-15.
Conover D.O. & Munch S.B. 2002. Sustaining fisheries
yields over evolutionary time scales. Science,
297: 94-96.
Copp G.H. 1989. The habitat diversity and fish reproduc-
tive function of floodplain ecosystems.
Environmental Biology of Fishes, 26: 1-27.
Crowl T.A., McDowell W.H., Covish A.P. & Johnson S.L.
2001. Freshwater shrimp effects on detrital
processing and nutrients in a tropical headwa-
ter stream. Ecology, 82: 775-783.
Dettmers J.M., Wahl D.H., Soluk D.A. & Gutreuter S.
2001. Life in the fast lane: Fish and foodweb
structure in the main channel of large rivers.
Journal of the North American Benthological
Society, 20: 255-265.
Drenner R.W., Smith J.D. & Threlkeld S.T. 1996. Lake
trophic state and the limnological effects of
omnivorous fish. Hydrobiologia, 319: 213-
223.
Dudgeon D. 2000. Large-scale hydrological changes in
tropical Asia: Prospects for riverine biodiversi-
ty. BioScience, 50: 793-806.
Fernandes C.C. 1997. Lateral migration of fishes in
Amazon floodplains. Ecology of Freshwater
Fish, 6: 36-44.
Flecker A.S. 1996. Ecosystem engineering by a dominant
detritivore in a diverse tropical stream.
Ecology, 77: 1845-1854.
Flecker A.S, Feifarek B.P. & Taylor B.W. 1999.
Ecosystem engineering by a tropical tadpole:
Density dependent effects on habitat structure
and larval growth rates. Copeia 1999: 495-
500.
Forsberg B.R., Araujo-Lima C.A.R.M., Martinelli L.A.,
Victoria R.L. & Bonassi J.A. 1993.
Autotrophic carbon sources for fish of the
Central Amazon. Ecology, 74: 643-652.
Galat D.L. & Zweimüller I. 2001. Conserving large-river
fishes: Is the highway analogy an appropriate
paradigm. Journal of the North American
Benthological Society, 20: 266-279.
Gomes L.C. & Agostinho A.A. 1997. Influence of the
flooding regime on the nutritional state and
juvenile recruitment of the curimba,
Prochilodus scrofa, Steindachner, in upper
Paraná River, Brazil. Fisheries Management
and Ecology, 4: 263-274.
304 Floodplain river food webs: Generalizations

Goulding M. 1980. The Fishes and the Forest. Berkeley,
CL, USA, University of California Press.
Goulding M.M., Carvalho L. & Ferreira E.G. 1988. Río
Negro: Rich life in poor water. The Hague:
SPB Academic Publishing.
Hahn N.S., de Andrian I.F., Fugi R. & Lescano de
Almeida V.L. 1997. Ecologia trófica. In
Editora da Universidade Estadual de Maringá.
A Planície de indundação do Alto Rio Paraná:
Aspectos Físicos, Biológicos e Socieconômicos.
Maringá. pp. 209-248.
Hall S.J. & Rafaelli D.G. 1993. Food webs: Theory and
reality. Advances in Ecological Research, 24:
187-239.
Hamilton S.K., Lewis W.M. Jr. & Sippel S.J. 1992.
Energy sources for aquatic animals in the
Orinoco River floodplain: Evidence from sta-
ble isotopes. Oecologia, 89: 324-330.
Handlos W.L. & Williams G.J. 1985. The Zambian sec-
tion of the Zambezi catchment– an overview.
In W.L. Handlos & G.W. Howard eds.
Development Prospects for the Zambezi Valley
in Zambia. Lusaka, Zambia, Kafue Basin
Research Committee, University of Zambia.
pp. 1-28.
Hein T., Heiler G., Pennetzdorfer D., Riedler P., Schagerl
M. & Schiemer F. 1999. The Danube restora-
tion project: Functional aspects and planktonic
productivity in the floodplain ecosystem.
Regulated Rivers: Research & Management,
15: 259-270.
Hocutt C.H. & Johnson P.N. 2001. Fish response to the
annual flooding regime in the Kavango River
along the Angola/Namibia border. South
African Journal of Marine Science, 23: 449-
464.
Holt R.D. 1996. Food webs in space: An island biogeo-
graphic perspective. In G.A. Polis & K.O.
Winemiller eds. Food Webs: Integration of
Patterns and Dynamics. New York, Chapman
& Hall. pp. 313-323.
Humphries P., King A.J. & Koehn J.D. 1999. Fish, flows
and flood plains: Links between freshwater
fishes and their environment in the Murray-
Darling River system, Australia. Environmen-
tal Biology of Fishes, 56: 129-151.
Humphries P., Luciano G.S. & King A.J. 2002. River reg-
ulation and fish larvae: Variation through space
and time. Freshwater Biology, 47: 1307-1331.
Huxel G.R. & McCann K. 1998. Food web stability: The
influence of trophic flows across habitats.
American Naturalist, 152: 460-469.
Jackson P.B.N. 1961. The impact of predation, especially
by the tiger-fish Hydrocynus vittatus Cast. on
African freshwater fishes. Proceedings of the
Zoological Society of London, 136: 603-622.
Jepsen D.B. & Winemiller K.O. 2002. Structure of tropi-
cal river food webs revealed by stable isotope
ratios. Oikos, 96: 46-55.
Jepsen D.B., Winemiller K.O. & Taphorn D.C. 1997.
Temporal patterns of resource partitioning
among Cichla species in a Venezuelan black-
water river. Journal of Fish Biology, 51: 1085-
1108.
Jepsen D.B., Winemiller K.O., Taphorn D.C. &
Rodriguéz-Olarte D. 1999. Variation in age
structure and growth of peacock cichlids from
rivers and reservoirs of Venezuela. Journal of
Fish Biology, 55: 433-450.
Junk W.J. 1989. Flood tolerance and tree distribution in
central Amazonian floodplains. In L.B. Holm-
Nielsen, I.C. Nielsen & H. Balslev eds.
Tropical Forests: Botanical Dynamics,
Speciation and Diversity. London, Academic
Press. pp. 47-64.
Junk W.J., Bayley P.B. & Sparks R.E. 1989. The flood
pulse concept in river-floodplain systems. In
D.P. Dodge ed. Proceedings of the internation-
al large rivers symposium. Can. J. Fish. Aquat.
Sci. Spec. Publ., 106: 110-127.
and implications for fisheries management 305

Kaufman L. 1992. Catastrophic change in species-rich
freshwater ecosystems. Bioscience, 42: 846-
858.
Kellerhals R. & Church M. 1989. The morphology of
large rivers: Characterization and manage-
ment. In D.P. Dodge ed. Proceedings of the
international large rivers symposium. Can. J.
Fish. Aquat. Sci. Spec. Publ., 106: 31-48.
Kline T.C. Jr., Goering J.J., Mathisen O.L., Poe P.H. &
Parker P.L. 1990. Recycling of elements trans-
ported upstream by runs of Pacific Salmon: I.
_15N and _13C evidence in Sashin Creek,
southeastern Alaska. Canandian Journal of
Fisheries and Aquatic Sciences, 47, 136-144.
Kramer D.L., Lindsey C.C., Moodie G.E.E. & Stevens
eds. 1978. The fishes and the aquatic environ-
ment of the central Amazon Basin, with partic-
ular reference to respiratory patterns.
Canadian Journal of Zoology, 56: 717-729.
Leite R.G., Araujo-Lima C.A.R.M., Victoria R.L. &
Martinelli L.A. 2002. Stable isotope analysis
of energy sources for larvae of eight fish
species from the Amazon floodplain. Ecology
of Freshwater Fish, 11: 56-63.
Lewis W.M., Jr., Hamilton S.K., Lasi M.A., Rodríguez
M. & Saunders J.F. III. 2000. Ecological deter-
minism on the Orinoco floodplain. BioScience,
50: 681-692.
Lewis W.M., Jr., Hamilton S.K., Rodríguez M., Saunders,
J.F. III & Lasi M.A. 2001. Foodweb analysis of
the Orinoco floodplain based on production
estimates and stable isotope data. Journal of
the North American Benthologial Society, 20:
241-254.
Loubens G. & Panfili J. 2001. Biologie de Piaractus
brachypomus Teleostei:Serrasalmidae dans
basin du Mamoré Amazonie bolivienne.
Ichthyological Explorations of Freshwaters,
12: 51-64.
Lowe-McConnell R.H. 1964. The fishes of the Rupununi
Savanna District of British Guiana, South
America. Pt I Ecological groupings of fish
species and effects of the seasonal cycle on the
fish. Journal of the Linnean Society Zoology,
45: 103-144.
Lowe-McConnell R.H. 1987. Ecological Studies of
Tropical Fish Communities. Cambridge, UK,
Cambridge University Press.
Lundberg J.G., Lewis W.M. Jr., Saunders J.F. III & Mago-
Leccia F. 1987. A major food web component
in the Orinoco River channel: Evidence from
planktivorous electric fishes. Science, 237: 81-
83.
Madsen T. & Shine R. 2000. Rain, fish and snakes:
Climatically driven population dynamics of
Arafura filesnakes in tropical Australia.
Oecologia, 124, 208-215.
Marrero C. 1987. Notas preliminares acerca de la historia
natural de los peces del Bajo Llano. Revista
Hydrobiología Tropical, 20: 57-63.
Melack J.M., Forsberg B.R.,Victoria R.L. & Richey J.E.
1999. Biogeochemistry of Amazon floodplain
lakes and associated wetlands. In M. McClain,
R. Victoria & J. Richey eds. The
Biogeochemistry of the Amazon Basin and Its
Role in a Changing World. Oxford, UK,
Oxford University Press. pp. 1-50.
Marchese M. & Ezcurra de Drago I.E. 1992. Benthos of
the lotic environments in the middle Parana
River system: Transverse zonation.
Hydrobiologia, 237: 1-13.
Oberdorff T., Guégan J.F. & Hugueny B. 1995. Global
scale patterns of fish species richness in rivers.
Ecography, 18: 345-352.
Pauly D., Christensen V., Dalsgaard J., Froese R. &
Torres F. Jr. 1998. Fishing down marine food
webs. Science, 279: 860-863.
Peterson C.C. 1997. Food webs of two Venezuelan clear-
water streams with seasonal fluctuations in
hydrology. College Station, TX, USA, A&M
University. (Master thesis)
306 Floodplain river food webs: Generalizations

Polis G.A. & Winemiller K.O. 1996. Food Webs:
Integration of Patterns and Dynamics. New
York, Chapman and Hall.
Polis G.A., Anderson W.B. & Holt R.D. 1997. Toward an
integration of landscape and food web ecolo-
gy: The dynamics of spatially subsidized food
webs. Annual Review of Ecology and
Systematics, 28: 289-316.
Power M.E. 1990. Effects of fish in river food webs.
Science, 250: 811-814.
Power M.E. 1992. Habitat heterogeneity and the func-
tional significance of fish in river food webs.
Ecology, 73, 1675-1688.
Power M.E., Sun A., Parker G., Dietrich W.E. & Wootton
T.J. 1995. Hydraulic food-chain models.
BioScience, 45: 159-267.
Pringle C.M., Freeman M.C. & Freeman B.J. 2000.
Regional effects of hydrological alterations on
riverine macrobiota in the New World:
Tropical-temperate comparisons. BioScience,
50: 807-823.
Putz R. & Junk W. 1997. Phytoplankton and primary pro-
duction. In The Central Amazon Floodplain.
Heidelberg, Springer-Verlag. pp. 207-222.
Quensiere J., Olivry J.-C., Poncet Y. & Wuillot J. 1994.
Environnment deltaïque. In J. Quensiere ed. La
Pêche dans le Delta Central du Niger. Paris:
ORSTOM Editions. pp. 29-80.
Rai H. & Hill G. 1984. Primary production in the
Amazonian aquatic ecosystem. In H. Sioli ed.
The Amazon. Limnology and Landscape
Ecology of a Mighty Tropical River and Its
Basin. Dordrecht, Dr. W. Junk Publishers. pp.
311-335.
Reinert T.R. & Winter K.A. 2001. Sustainability of har-
vested pacú Colossoma macropomum popula-
tions in the northeastern Bolivian Amazon.
Conservation Biology, 16: 1344-1351.
Remsen J.V. & Parker T.A. 1983. Contribution of river-
created habitats to bird species richness in
Amazonia. Biotropica, 15: 223-231.
Resh V.H., Hildrew A.G., Statzner B. & Townsend C.R.
1994. Theoretical habitat templets, species
traits, and species richness: A synthesis of
long-term ecological research on the Upper
Rhone River in the context of concurrently
developed ecological theory. Freshwater
Biology, 31: 539-554.
Ribeiro M.C.L. de B. & Petrere M. Jr. 1990. Fisheries
ecology and management of the jaraqui
Semaprochilodus taeniurus, S. insignis in cen-
tral Amazonia. Regulated Rivers: Research &
Management, 5: 195-215.
Robinson C.T. Tockner K. & Ward J.V. 2002. The fauna
of dynamic riverine landscapes. Freshwater
Biology, 47: 661-677.
Rodríguez M.A. & Lewis, W.M. Jr. 1997. Structure of
fish assemblages along environmental gradi-
ents in floodplain lakes of the Orinoco River.
Ecological Monographs, 67: 109-128.
Schiemer F. 1999. Conservation of biodiversity in flood-
plain rivers. Archives fur Hydrobiologie
Supplement 115/3. pp. 423-438.
Schiemer F., Keckeis H., Winkler G. & Flore L. 2001a.
Large rivers: The relevance of ecotonal struc-
ture and hydrological properties for the fish
fauna. Archives fur Hydrobiologie Supplement
135/2-4. pp. 487-508.
Schiemer F., Keckeis H., Reckendorfer W. & Winkler G.
2001b. The inshore retention concept and its
significance for large rivers. Archives fur
Hydrobiologie, 135/2: 509-516.
Schmid-Araya J.M. & Schmid P.E. 2000. Trophic rela-
tionships: Integrating meiofauna into a realis-
tic benthic food web. Freshwater Biology, 44:
149-163.
and implications for fisheries management 307

Sedell J.R., Reeves G.H., Hauer F.R., Stanford J.A. &
Hawkins C.P. 1990. Role of refugia in recov-
ery from disturbances: Modern fragmented
and disconnected river systems. Environmen-
tal Management, 14: 711-724.
Sheppe W. & Osborne T. 1971. Patterns of use of a flood
plain by Zambian mammals. Ecological
Monographs, 41: 179-207.
Shiel R.J., Green J.D. & Nielsen D.L. 1998. Floodplain
biodiversity: Why are there so many species?
Hydrobiologia, 387/388: 39-46.
Smock L.A. 1994. Movements of invertebrates between
stream channels and forested floodplains.
Journal of the North American Benthological
Society, 13: 524-531.
Southwood T.R.E. 1977. Habitat, the templet for ecolog-
ical strategies? Journal of Animal Ecology, 46:
337-365.
Sparks R.E. 1995. Need for ecosystem management of
large rivers and their floodplains. BioScience,
45: 168-182.
Thorp J.H. & Delong M.D. 1994. The riverine productiv-
ity model: An heuristic view of carbon sources
and organic processing in large river ecosys-
tems. Oikos, 70: 302-308.
Thorp J.H. & Delong M.D. 2002. Dominance of
autochthonous autotrophic carbon in food
webs of heterotrophic rivers. Oikos, 96: 543-
550.
Thorp J.H., Delong M.D., Greenwood K.S. & Casper A.F.
1998. Isotopic analysis of three food web the-
ories in constricted and floodplain regions of a
large river. Oecologia, 117: 551-563.
Tilman D. 1999. The ecological consequences of changes
in biodiversity: The search for general princi-
ples. Ecology, 80: 1455-1474.
Tockner K., Pennetzdorfer D., Reiner N. Schiemer F. &
Ward J.V. 1999. Hydrological connectivity and
the exchange of organic matter and nutrients in
a dynamic river-floodplain system. Freshwater
Biology, 41: 521-535.
Toth L.A., Melvin S.L., Arrington D.A. & Chaimberlain
J. 1998. Hydrologic manipulation of the chan-
nelized Kissimmee River. Bioscience, 48: 757-
764.
Tucker J.K., Theiling C.H. & Camerer J.B. 1996.
Utilization of backwater habitats by unionid
mussels Bivalvia: Unionidae on the lower
Illinois River and in pool 26 of the upper
Mississippi River. Transactions of the Illinois
Academy of Science, 89: 113-122.
Van den Brink F.W.B., de Leeuw J.P.H.M, van der Velde
G. & Verheggen G.M. 1993. Impact of hydrol-
ogy on the chemistry and phytoplankton devel-
opment in floodplain lakes along the lower
Rhine and Meuse. Biogeochemistry Dordrecht,
19: 103-128.
Vanni M.J. 1996. Nutrient transport and recycling by con-
sumers in lake food webs: Implications for
plankton. In G.A. Polis & K.O. Winemiller eds.
Food Webs: Integration of Patterns and
Dynamics. New York, Chapman and Hall. pp.
81-95.
de Vazzoler A.E.M., Amadio S.A. & Daraciolo-Malta
M.C. 1989. Aspectos biológicos de peixes
Amazônicos. XI. Reprodução das espécies do
gênero Semaprochilodus Characiformes,
Prochilodontidae no baixo Rio Negro,
Amazonas, Brasil. Revista Brasiliera de
Biologia, 49: 165-173.
Wantzen K.M., Machado F.D.A., Voss M., Boriss H. &
Junk W.J. 2002. Seasonal isotopic shifts in fish
of the Pantanal wetland, Brazil. Aquatic
Science, 64: 239-251.
Ward J.V. & J.A. Stanford. 1989. Riverine ecosystems:
The influence of man on catchment dynamics
and fish ecology. In D.P. Dodge ed.
Proceedings of the international large rivers
symposium. Can. J. Fish. Aquat. Sci. Spec.
Publ., 106: 56-64.
308 Floodplain river food webs: Generalizations

Ward J.V., Tockner K. & Schiemer F. 1999. Biodiversity
of floodplain river ecosystems: Ecotones and
connectivity. Regulated Rivers: Research &
Management 15: 125-139.
Welcomme R.L. 1985. River Fisheries. Food and
Agriculture Organization Fisheries Technical
Paper 262. Rome, FAO. pp. 1-330.
Welcomme R.L. 1990. Status of fisheries in South
American Rivers. Interciencia, 15: 337-345.
Willson M.F., Gende S.M. & Marston B.H. 1998. Fishes
and the forest: Expanding perspectives on fish-
wildlife interactions. Bioscience, 48: 455-462.
Wilson J.A., French J., Kleban P., McKay S.R. &
Townsend R. 1991. Chaotic dynamics in a
multispecies fishery: A model of community
predation. Ecological Modeling, 58: 303-322.
Winemiller K.O. 1989a. Ontogenetic diet shifts and
resource partitioning among piscivorous fishes
in the Venezuelan llanos. Environmental
Biology of Fishes, 26: 177-199.
Winemiller K.O. 1989b. Patterns of variation in life his-
tory among South American fishes in seasonal
environments. Oecologia, 81: 225-241.
Winemiller K.O. 1990. Spatial and temporal variation in
tropical fish trophic networks. Ecological
Monographs, 60: 331-367.
Winemiller K.O. 1991a. Ecomorphological diversifica-
tion of freshwater fish assemblages from five
biotic regions. Ecological Monographs, 61:
343-365.
Winemiller K.O. 1991b. Comparative ecology of
Serranochromis species Teleostei: Cichlidae in
the Upper Zambezi River. Journal of Fish
Biology, 39: 617-639.
Winemiller K.O. 1996a. Factors driving spatial and tem-
poral variation in aquatic floodplain food
webs. In G.A. Polis & K.O. Winemiller eds.
Food Webs: Integration of Patterns and
Dynamics. New York, Chapman and Hall. pp.
298-312.
Winemiller K.O. 1996b. Dynamic diversity: Fish com-
munities of tropical rivers. In M.L. Cody &
J.A. Smallwood eds. Long-term Studies of
Vertebrate Communities. Orlando, FL, USA,
Academic Press. pp. 99-134.
Winemiller K.O. & Jepsen D.B. 1998. Effects of season-
ality and fish movement on tropical river food
webs. Journal of Fish Biology, 53 Supplement
A. pp. 267-296.
Winemiller K.O. & Jepsen D.B. 2002. Migratory
neotropical fish subsidize food webs of olig-
otrophic blackwater rivers. In G.A. Polis, M.E.
Power & G. Huxel eds. Food Webs at the
Landscape Level, Chicago, University of
Chicago Press. (forthcoming)
Winemiller K.O. & Rose K.A. 1992. Patterns of life-his-
tory diversification in North American fishes:
Implications for population regulation.
Canadian Journal of Fisheries and Aquatic
Sciences, 49: 2196-2218.
Winemiller K.O., Marrero C. & D.C. Taphorn D.C. 1996.
Perturbaciones causadas por el hombre a las
poblaciones de peces de los llanos y del piede-
monte Andino de Venezuela. Biollania
Venezuela, 12: 13-48.
Winemiller K.O., Tarim S., Shormann D. & Cotner J.B.
2000. Spatial variation in fish assemblages of
Brazos River oxbow lakes. Transactions of the
American Fisheries Society, 129, 451-468.
Woodward G. & Hildrew A.G. 2002. Food web structure
in riverine landscapes. Freshwater Biology, 47:
777-798.
Zaret T.M. & Paine R.T. 1973. Species introduction in a
tropical lake. Science, 182: 449-455.
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