J. Phycol. 34, 741–749 (1998)
MINIREVIEW
USE OF PHYTOPLANKTON IN LARGE RIVER MANAGEMENT1
John D. Wehr2
Louis Calder Center-Biological Station, Fordham University, P.O. Box K, Armonk, New York 10504
and
Jean-Pierre Descy
Facultés Universitaires Notre-Dame de la Paix, Unit of Freshwater Ecology, 61 Rue de Bruxelles, B-5000 Namur, Belgium
some authors define ‘‘large river’’ as a fluvial system
large enough to intimidate researchers (Hynes
1989) or too large to study without a boat and specialized equipment (Stalnaker et al. 1989). Streams
and rivers are classified by geomorphologists according to fluvial and watershed properties, which recognize a continuum of physical features (channel
width, average depth, discharge, catchment area) as
one progresses downstream (Morisawa 1968). When
viewed from this perspective, the distinction between large and smaller rivers may be considered
arbitrary (Stalnaker et al. 1989, Townsend 1996).
Ecological theory developed for river ecosystems,
such as the RCC (Vannote et al. 1980) and Nutrient
Spiraling model (Newbold et al. 1981), employ longitudinally integrated processes as factors affecting
resource supply and ecosystem function. The relevance and usefulness of the RCC for large rivers has
been questioned (Sedell et al. 1989, Thorp and Delong 1994, Townsend 1996), but the definition of
‘‘large river’’ that has emerged from these studies is
that of rivers greater than sixth order (based on
Strahler 1957). Other workers examining the importance of phytoplankton production to carbon
supply and system function have suggested that simply ‘‘low-gradient’’ or ‘‘lowland’’ rivers may be more
useful concepts because their metabolism is frequently dominated by algal or other in situ autotrophic organisms, despite the fact that many are lightlimited due to depth and turbidity (Baker and Baker
1979, Descy and Gosselain 1994, Köhler 1995, Reynolds and Descy 1996). This latter perspective has
merit, because downstream reaches of smaller rivers
possess many features common to large rivers
(floodplain influences, high turbidity, lateral carbon
inputs, lower gradient). Other authors maintain that
large rivers should be viewed as systems distinct
from upland sections, with ecological features and
management problems that deserve particular attention (Sedell et al. 1989, Admiraal et al. 1994). Indeed, large rivers are widely recognized and even
defined as manipulated or regulated ecosystems
(Ward and Stanford 1983, Décamps 1996). Therefore, our review focuses on management concerns
Key index words: freshwater; large rivers; phytoplankton;
river impoundments
Abbreviations: POC, particulate organic carbon; RCC, river continuum concept; SRP, soluble reactive phosphate
Historically, rivers have served as sources of drinking water, fisheries resources, transportation routes,
irrigation supplies, and waste removal systems. Human civilization has had many major effects on rivers, dating back more than 5000 years when Egyptians built dams on the Nile to supply water for
crops and human consumption. Today, management of large rivers requires a balance between human needs and ecological integrity, although until
quite recently, ecological principles have played a
minor role in river management (Edwards 1995).
Planktonic algae are an important part of these
issues because they play a central role in the functioning of large rivers. Algal communities are major
producers of organic carbon in larger rivers, are a
food source for planktonic consumers, and may represent the primary oxygen source in many low-gradient rivers (Thorp and Delong 1994, Köhler 1995,
Reynolds and Descy 1996). Phytoplankton are responsive to excessive supplies of inorganic nutrients
and may pose problems in long stretches of rivers
with cultural eutrophication, but may also enhance
water quality for humans in rivers affected by agricultural or industrial uses. Algal communities of river systems consist not only of suspended algae, but
also a diverse benthic assemblage of macrophytic
forms, smaller epilithic species, epiphytes, and sediment-dwelling forms (Reynolds 1996). This review
focuses on the use of planktonic algae in river management because of their central importance in larger rivers and because of the growing need for ecosystem-level studies on river plankton.
WHAT ARE LARGE RIVERS?
Ecologists know much more about small tributaries than larger rivers, in large part because they are
simply easier to study. Undoubtedly, this is why
1
Received 5 November 1997. Accepted 8 June 1998.
Author for reprint requests;
e-mail: [email protected].
2
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JOHN D. WEHR AND JEAN-PIERRE DESCY
FIG. 1. Schematic representation of alterations and disturbances
that affect ecosystem integrity in large rivers. A management
scheme may alter the number and position of sampling stations
in an attempt to monitor specific effects, but due to the directional flow of rivers, each location will represent a combination
(and perhaps interaction of) factors from further upstream.
of lowland rivers greater than sixth order and the
role that phytoplankton may play in these systems.
MANAGEMENT ISSUES
Many studies document the effects of human disturbances on communities in large and small rivers,
but few of these findings have been adequately communicated to water management agencies (Petts et
al. 1995). This condition exists despite the fact that
most large rivers of the world are now or will soon
be substantially altered by human activities (Sparks
1995). A few of the important disturbances include
channelization, navigation or hydroelectric dams,
removal of adjacent wetlands, toxic pollutants, and
cultural eutrophication (Johnson et al. 1995, Petts
et al. 1995). Management of these large-river disturbances is complicated by the fundamental feature
common to all flowing waters: activities or disturbances at one location affect processes and organisms downstream (Fig. 1). Due to the longitudinal
nature of rivers, management problems that arise
are frequently more difficult to manage than in
lakes. For example, excessive nutrients may enter
the system from several points along the main channel, and the forms of nutrients also depend on various watershed processes (e.g. hydrogeological processes as nutrients transfer through soils and
groundwater), which are not well understood.
Downstream transport of dissolved and particulate
substances often requires that management schemes
consider large river stretches (.102–103 km) or entire basins, as well as adjacent drainage areas.
Human disturbances affect not only riverine organisms and processes, but also the links between a
river and its watershed, by disrupting lateral linkages
with floodplains and groundwater (Townsend
1996). Perhaps this is one reason why computer
models that have attempted to predict phytoplankton production or species composition in large rivers have met with only modest success (Billen et al.
1994, Ruse and Love 1997). River ecosystems are
actually networks of interactions among physical,
chemical, and biological processes, which reach a
higher degree of complexity downstream. There are
also longitudinal differences in the time scales of
chemical and biological processes. Such complexity
renders it difficult to design policies and assess the
results of management actions. Recently, studies
have taken into account riverine, watershed, and
land use variables, as well as differences in spatial
scale in devising management models for river water
and fish habitat quality (Hunsaker and Levine 1995,
Roth et al. 1995, Smitz et al. 1997).
Use of phytoplankton in management has had a
fairly long history. Concern most often centers on
adverse effects, such as taste and odor problems, filter clogging at water treatment plants, and effects
on fisheries (Palmer 1962). Fewer studies have documented the beneficial uses of algae in management practice. How algae are applied depends on
the problem, as well as on river size and flow regime.
For example, in the highly regulated Buffalo River
(South Africa), high nutrient loads were reduced by
sedimenting algal populations along impounded
stretches, resulting in the ability to release relatively
low-nutrient water below the dams (O’Keefe et al.
1990). Studies using algal community structure or
accumulation of toxic metals have been employed
in large-scale analyses of water quality in many
streams and rivers (del Giorgio et al. 1991, Whitton
et al. 1991, Lowe and Pan 1996). An important complication to their use in management is that the
composition, dynamics, and production of algal
communities in large rivers is less well understood
than in either lakes or small streams.
RIVER PHYTOPLANKTON ECOLOGY: A SHORT REVIEW
Various aspects of potamoplankton, the community of suspended algae in flowing water, have been
reviewed by Reynolds and Descy (1996). These authors highlight several key points specific to the
composition and ecology of planktonic algal assemblages in large rivers. Ironically, many of the most
important features emphasized were understood
long ago (Kofoid 1908, Krieger 1927, Reinhard
1931, Butcher 1932, Welch 1952). Among questions
left unresolved, the origin of phytoplankton assemblages in running waters is most often raised. Even
the earliest studies perceived that algal inocula from
various sources contributed to the main channel,
but potamoplankton is also ‘‘native’’ to rivers, as
demonstrated more precisely by Reynolds (1995)
and Reynolds and Glaister (1993). The actual
sources of inocula (backwaters connected to the
main channel or dead zones within the channel) are
still uncertain and likely vary among river systems.
It is tempting to speculate that most planktonic algae in rivers are meroplanktonic (planktonic with a
benthic life stage) and recolonize littoral areas or
the river bed for survival and perennation. The benthos clearly serves as the origin of suspended algae
in smaller streams (Swanson and Bachmann 1976).
PHYTOPLANKTON AND LARGE RIVER MANAGEMENT
In such cases, the chlorophyll a flux is positively related to discharge. However in large rivers, phytoplankton biomass varies inversely with discharge
(Schmidt 1994); this distinction is an important
characteristic that separates small from large rivers.
Composition and selection. The most successful algal
groups in large rivers are diatoms and green algae,
the latter tending to dominate in summer conditions, although this pattern is far from consistent.
Cryptophytes and cyanobacteria may also at times be
abundant, whereas other groups, chrysophytes, dinoflagellates, and euglenophytes are usually less developed, despite records of blooms of particular taxa
in specific conditions (e.g. Ohio River: Wehr and
Thorp 1997, St. Lawrence River: Hudon et al. 1996).
Plankton surveys provide lists of up to several hundred taxa, but the list of common and truly planktonic taxa is much less (Reynolds and Descy 1996),
and (unlike benthic macroalgal species), few if any
planktonic taxa are specific to rivers (Round 1981).
The principal selective factors affecting river phytoplankton were clearly summarized by Reynolds
(1988, 1995):
● Nutrient limitation is unlikely in rivers, despite
the influence of algal growth on nutrient (Si, P)
levels. Nutrients in rivers are often in considerable excess of algal requirements, such that models based on nutrient resource ratios in lakes (Tilman et al. 1982) are not applicable in most rivers.
● Phytoplankton abundance and production are
controlled by discharge. This is related to residence time, channel depth, and dilution rate and
affects water transparency and sedimentation.
● Maintaining stable or persistent phytoplankton
communities is impossible in large (deep and turbid) rivers if these systems exist as fully mixed
reaches. Therefore, river mixing must be incomplete (Margalef 1960), which is provided by heterogeneity in channel morphology.
The river environment selects for fast-growing
species able to cope with wide variations in light conditions, which depend on incident light, water transparency, and depth (not unlike turbid, well-mixed
shallow lakes; Reynolds et al. 1994). There are very
few studies demonstrating nutrient limitation in
larger rivers (Moss and Balls 1989) or dependence
of algal growth upon P availability (Reynolds and
Descy [1996], but see Basu and Pick [1996] for a
contrary view). Such issues are important when dealing with the effects of eutrophication in rivers. Analysis of long-term data shows that increases in algal
biomass that have occurred in recent decades in
many European and North American rivers may
have resulted from increased nutrient inputs coupled with hydrological changes and river regulation.
Biomass, longitudinal variation, and driving variables.
A wide range of phytoplankton biomass has been
reported for rivers, from ,1 mg (Orinoco River, Brazil: Lewis 1988) up to ø400 mg chlorophyll a·L21
743
(several rivers: Jones 1984, Moss et al. 1984). The
largest values are comparable to those reported for
highly eutrophic ponds. It is in the higher range
where potamoplankton may cause harmful effects
for water use and even decreases in oxygen concentrations. However, as algal photosynthesis is often
the major supplier of oxygen to slowly flowing rivers,
phytoplankton loss may also result in water quality
problems (Descy 1992).
That potamoplankton biomass exhibits variations
along the course of a river has been known for many
decades (Welch 1952). Simulation models show this
pattern develops in four phases, progressing downstream (no plankton in the headwaters, increase,
maximum, decline), and results from interactions
among river morphology, hydrology, light availability, and algal growth rate (Descy and Gosselain 1994,
Garnier et al. 1995). Growth rates in eutrophic rivers strongly depend on the balance between algal
photosynthesis and respiration, which is positive in
shallow, transparent river channels and may become
negative in deep and turbid reaches. This deficit in
phytoplankton growth is particularly true in river estuaries, such as the tidal Hudson River, New York
(Cole et al. 1992). Phytoplankton in lower reaches
are likely imported from upstream (less turbid)
stretches that support positive net algal production.
These longitudinal dynamics are driven by downstream transport of plankton into reaches where
growth fluctuates in response to physical factors;
therefore, measures against eutrophication must
consider the entire river system.
The view that potamoplankton development in
large rivers takes place mostly in the main channel
and is heavily dependent on physical characteristics
is applicable not just to regulated or eutrophic rivers. The same processes may govern phytoplankton
growth in rivers with extensive floodplains, as shown
by Lewis (1988) for the Orinoco River. On an annual basis, plankton gross production was 9.5 g
C·m 2 2 ; the floodplain contributed only 0.25 g
C·m22. At the upper end of the production range,
the River Loire, France (Lair and Sargos 1993, Lair
and Reyes-Marchant 1997), provides an example of
a river with massive potamoplankton growth, but
limited flow regulation, where exchanges (nutrients,
organisms) with the littoral zone substantially enhance plankton development.
Despite many studies to the contrary, nutrient limitation should not be excluded as a factor in the
longitudinal development of river phytoplankton.
Data from rivers in southeastern Ontario and western Quebec reveal a positive relationship between
total P and chlorophyll a (Basu and Pick 1996). We
suggest, however, that a clear demonstration of nutrient limitation in river systems is generally still lacking. Such a relation is likely true in systems with few
nutrient sources, long residence times, and high water clarity (or shallow depth). Few large rivers worldwide combine these conditions, and many if not
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JOHN D. WEHR AND JEAN-PIERRE DESCY
most are already significantly affected by human influences.
Role of loss processes in river phytoplankton dynamics.
Phytoplankton development in some rivers may be
favored because of lower grazing pressure and reduced sedimentation rates (Cole et al. 1991, 1992).
There are some data indicating that zooplankton
biomass in lotic environments is lower than in lakes
(Pace et al. 1992, Thorp et al. 1994) and that phytoplankton–zooplankton interactions are weak in
most river systems (Köhler 1995, Basu and Pick
1996). However, declines in potamoplankton numbers or biomass do occur in otherwise favorable conditions for algal growth (de Ruyter van Steveninck
et al. 1992, Gosselain et al. 1994). These declines
may result from poor light availability combined
with zooplankton grazing and sedimentation losses.
However, few field studies have quantified the impact of zooplankton grazing on phytoplankton in
large rivers (Thorp et al. 1994, Gosselain et al.
1998). The primary constraint in rivers is residence
time, which affects both phytoplankton and their
grazers (Pace et al. 1992, Thorp et al. 1994). Rivers
usually select for small-bodied zooplankton with an
ability to grow rapidly enough to compensate for
limited residence times (Viroux 1997), but which
may have a low capacity to regulate phytoplankton
biomass. Several authors have reported decreases in
phytoplankton numbers brought about by benthic
filter feeders, especially zebra mussels (Dreissena polymorpha: Effler et al. 1996, Roditi et al. 1996, Strayer
et al. 1996) and other invaders, such as Corbicula sp.
and Corophium curvispinum (Bachmann et al. 1995).
Voracious feeding rates coupled with very large populations have dramatic effects on phytoplankton, water quality, and ecosystem processes in large rivers.
These recent and dramatic developments in many
large rivers (e.g. Mississippi, Ohio, St. Lawrence,
Hudson) call for specific management actions.
Almost no studies have shown that sedimentation
in rivers is an important loss process for phytoplankton. To date, sedimentation losses have been mostly
calculated, taking into account water velocity, depth,
and specific settling velocity of algal species (Carling
1992). However, it is likely that sedimentation is a
selective factor that may suppress diatom numbers
and perhaps prevent their dominance in shallow river stretches, especially in regions of very low flow
(Reynolds 1995).
CASE STUDIES AND OTHER MANAGEMENT ISSUES
Eutrophication in large rivers. Most large European
rivers have a long record of eutrophication attributed to human influences (Friedrich and Viehweg
1984, Descy 1987), which raises concern about the
sustainability of domestic uses of river water. This is
especially true when river water is abstracted and
intensively used, such as with the River Meuse (Belgium), a source of drinking water for about six million people (Descy 1992). Although increased algal
FIG. 2. Seasonal changes in maximum and 8- year average phytoplankton densities in the Hungarian section of the River Danube during 1957–1965 and 1979–1987. The 10-fold increase in
(nonwinter) algal densities occurred during a period of similar
nutrient levels, but marked changes in the hydrological conditions brought on by upstream reservoir construction (redrawn
after Kiss 1994).
growth is commonly thought of as resulting principally from increased nutrient loading, it is far from
clear that this is true in most rivers. For example, in
the case of River Danube (Hungary), there are excellent algological records spanning .40 years (Szemes 1967, Kiss 1994). A 10-fold increase in phytoplankton densities occurred during the 1970s (Fig.
2) without a noticeable increase in the nutrient supply since the end of the 1950s. Rather, changes in
suspended matter transport, brought about by reservoirs in the German and Austrian sections, improved water transparency and light availability for
algal growth. It is also likely that increased residence
time from the exploitation of the upper Danube favored the development of euplanktonic species in
the river.
A similar phytoplankton increase was observed in
the River Meuse, Belgium, by water supply companies since the early 1970s (Descy 1992). Strong increases in mean annual chlorophyll a occurred with
no clear trend in P concentrations within the Belgian sector (SRP: 50–100 mg·L21). Despite increases
in N concentrations in the Meuse (especially NO32),
PHYTOPLANKTON AND LARGE RIVER MANAGEMENT
as in other European rivers (Van Dijck 1996), levels
in the 1970s were already in the mg·L21 range, well
exceeding algal demand. Phytoplankton blooms in
the Meuse may have built up in the French sector
(Léglize and Salleron 1988) as a result of water
treatment plants lacking P-stripping and other tertiary treatment capabilities. The Meuse received increased P in river reaches that already had optimal
hydrological features for phytoplankton growth. As
a consequence, algal biomass represented an average of 58% of the total POC transported in the upper Belgian section of the Meuse; similar values have
been reported for other large rivers (Descy and Gosselain 1994).
Similar long-term, high nutrient levels (1960 to
present) and flow regulation have also prevailed in
the Ohio River (Seilheimer 1963, Nall 1965, Wehr
and Thorp 1997). Under these conditions, it is unclear whether hydraulic or chemical factors play the
major role in favoring high biomass. The upper Mississippi also supports large populations of algal species regarded as typical of eutrophic waters (Microcystis aeruginosa, Aphanizomenon flos- aquae, Aulacoseira granulata, A. italica), and although nutrient levels
were generally high, there was no evidence that
standing crop correlated with N or P concentration
(Lange and Rada 1993, Huff 1986). Authors also
contend that discharge and temperature were the
most likely controlling factors of algal biomass
through navigation dams and channelization in the
Mississippi. Such conditions are common in most
large rivers of the world and illustrate that even
when long-term nutrient data are available, it is not
always straightforward to make conclusions about
the primary causes of eutrophication.
Some eutrophic rivers carry such heavy burdens
of N and P that identification of the critical limiting
nutrient for regulation by a management agency
may not be possible. This is because standard nutrient addition bioassays (e.g. Cain and Trainor 1973,
Gerhart and Likens 1975) will not produce the necessary growth responses when supplies already exceed demand. An alternative method has been devised for the periodically hypereutrophic Neuse River (North Carolina), in which ‘‘dilution bioassays’’
apply stepwise dilutions of N, P, Fe, and trace metals
(Paerl and Bowles 1987). Such tests identify the nutrient(s) most responsible for triggering massive algal growth and help set target nutrient levels that
management agencies may recommend for river systems.
Solutions to reduce nuisance blooms in these very
eutrophic rivers may lie in hydraulic control (i.e.
natural flow conditions) whenever possible but must
also involve P reduction. At least a 90% reduction
in P is required in many large rivers because most
of the algal species that thrive in large rivers such
as in the Danube, Meuse, and Ohio (e.g. small species of Stephanodiscus, Cyclotella, and Chlorella-like
green algae) have half-saturation growth constants
745
#10 mg P·L21 (Van Donk and Kilham 1990, Lampert and Sommer 1997). This is ,10% of current P
concentrations in many rivers. Furthermore, taking
into account intracellular storage and downstream
displacement of water masses, careful treatment
strategies must be designed to induce P limitation
in large rivers. To develop and test such strategies,
reliable simulation models must quantify nutrient
inputs, river network features, and biological responses.
Potamoplankton occupies a key place in the oxygen budget of a river at least during the growing
season. On one hand, algal photosynthesis is a major source of oxygen production in most large and/
or lowland rivers. On the other hand, organic matter production in eutrophic rivers in excess of grazer
requirements is principally degraded by heterotrophic bacteria following phytoplankton mortality,
and can lead to severe oxygen depletion. These deleterious effects may be seen as a result of river-based
power plants, as demonstrated in model simulations
of the James River estuary, Virginia (Smith and Jensen 1974). These studies show how increased eutrophication may interfere with industrial uses of river
water, affecting algal mortality and physiological
changes from thermal shock. Further effects include
chlorine treatment and entrainment, the effects of
which are often difficult to sort out within the complex situations in the field. Finally, eutrophication
of large rivers discharging into the sea have also had
serious effects on water quality of coastal regions,
through increases in nutrient loading and changes
in nutrient ratios (Rabalais et al. 1996, Billen and
Garnier 1997, Lorenz et al. 1997).
Effects of exotic invaders. Recent literature has focused on phytoplankton declines resulting from biological invasions in large rivers, as well as lakes.
Such decreases in phytoplankton biomass have been
documented from several European and North
American rivers (de Ruyter van Steveninck et al.
1992, Descy 1993, Effler and Siegfried 1994, Gosselain et al. 1994, Garnier et al. 1995, Köhler 1995),
although not all have received a definitive explanation. Zooplankton grazing may play a role in some
cases (Gosselain et al. 1998), although short-lived
zooplankton that develop in rivers are generally unable to control algal biomass over longer periods of
time (i.e. several weeks). The most studied filter
feeders affecting phytoplankton declines is the zebra mussel Dreissena polymorpha, which invaded
North American rivers in the early 1990s (Effler et
al. 1996, Roditi et al. 1996, Strayer et al. 1996, Cope
et al. 1997, Gist et al. 1997). In the Hudson River,
large populations built up between 1992 and 1994
and reduced phytoplankton abundance by a factor
of 5 to 10 compared to preinvasion levels (Caraco
et al. 1997). Moreover, unlike small planktonic herbivores, zebra mussels apparently feed effectively on
most phytoplankton sizes, thereby controlling a
large fraction of riverine algal biomass. Similar ef-
746
JOHN D. WEHR AND JEAN-PIERRE DESCY
fects have been described in the Seneca River, New
York (Effler et al. 1996), where large densities of
mussels also contribute to low oxygen concentrations. Although perhaps less dramatic, some European rivers have suffered from recent invasions by
Dreissena and other exotic filter feeders (Bachmann
et al. 1995). What measures should be taken to reduce the spread of these organisms is not clear, and
there is a debate about whether they may be limited
by availability of substratum or food (Strayer et al.
1996). It is reasonable to assume, however, that pollution control measures to reduce algal proliferation
(e.g. reductions in P inputs) may have at least two
beneficial effects on these rivers: (1) decreases in
the intensity and frequency of nuisance blooms and
(2) a smaller food supply for zebra mussels. Some
predictions may be possible through simulation
models, provided that reliable data on filter feeder
densities, population dynamics, and in situ filtering
rates are available. But long-term algal monitoring
coupled with reductions in nutrient loading is still
necessary.
Cyanobacteria in rivers. A large contribution of colonial cyanobacteria to potamoplankton has been
most often reported from downstream reaches of
warm-temperate and tropical rivers, such as the Rio
Salado, Argentina (O’Farrell 1993), Nagdonk River,
South Korea (Ha et al. 1998), Neuse River, North
Carolina (Paerl and Bowles 1987), and Darling River, Australia (Hötzel and Croome 1994). However,
cyanobacteria blooms comprising potentially toxic
strains have been reported periodically in cold temperate rivers. These include the Ohio River over at
least three decades (Seilheimer 1963, Nall 1965, Petersen and Stevenson 1989, Wehr and Thorp 1997)
and the upper Mississippi since the 1930s (e.g. Reinhard 1931, Baker and Baker 1979, Huff 1986).
Clearly, cyanobacterial blooms in large rivers are not
a recent phenomenon: Microcystis blooms have been
reported since 1930 in the Potomac River, Maryland
(Krogmann et al. 1986). However, slow growth rates
of many of these organisms likely do not allow them
to develop large populations in large rivers unless
there are relatively high residence times. Accordingly, in the Rhine and Meuse (Germany and Belgium), bloom-forming cyanobacteria build large
populations in lower reaches of these rivers, but only
for short periods during summer (Ibelings et al.
1998). Interestingly, the 1995 survey in the Rhine
indicated that cyanobacteria developed large populations in summer in the Untersee, but quickly disappeared in the river downstream. A significant portion of the bloom (Planktothrix agardhii, plus small
Chroococcales) was later observed near Lobith,
ø800 km downstream of Untersee.
The conditions under which cyanobacterial
blooms develop in the Murray River (Australia) have
been studied in detail by Bormans et al. (1997). Although slow flow (current speed ,0.05 m·s21) has
been put forward as a key factor, temperature strat-
ification is also usually necessary for the development of Anabaena circinalis blooms. It is interesting
to note that the model used for simulating the Aulacoseira granulata–Anabaena sp. alternation in the
Murray River comprised a detailed physical submodel and a simple algal submodel based on algal
growth (as a function of light) and a buoyancy coefficient for the two taxa. Sufficient discharge seems
the best measure to suppress cyanobacterial blooms,
perhaps along with artificial destratification (Webster et al. 1996).
While the risk of potentially harmful cyanobacteria blooms in large temperate rivers is small, some
systems with high residence times or connected to
eutrophic ponds or lakes may still be prone to
bloom development (e.g. River Bure, Moss et al.
1984). Nonetheless, it is important to remain concerned with the possible expansion of cyanobacterial species (Couté et al. 1997, Padisák 1997) and to
be able to assess the risk of harmful blooms in fluvial
systems.
ARE POTAMOPLANKTON USEFUL INDICATORS OF THE
ECOLOGICAL STATUS OF RIVERS?
In lakes, phytoplankton biomass has been used
for decades to assess trophic status and to identify
artificial eutrophication induced by human activities
(Wetzel 1983, Harper 1992). Moreover, detailed
studies based on long-term records of phytoplankton numbers and taxonomic composition have
shown that, in general, phytoplankton are very sensitive indicators of various environmental changes
(Maberly et al. 1994). So far, no such biomass–nutrient indicator system has been successfully applied
in large or lowland rivers, mainly because longitudinal dynamics prevent the reliable use of simple
trophic scales based on chlorophyll a. Potamoplankton dynamics respond primarily to physical factors
and may fluctuate considerably in time and space.
Perhaps this is why phytoplankton abundance and
biomass data from large rivers frequently do not correlate strongly with nutrient chemistry, such as in
the River Severn, U.K. (Ruse and Love 1997), or
may even correlate negatively, as in the Ohio River
(Wehr and Thorp 1997).
Physical factors, along with nutrients and biotic
interactions (grazing), also influence community
composition, but we are a long way from being able
to sort out which are the most important determinants at the species level. Even at the level of major
taxonomic groups, only basic trends can be identified (Reynolds 1994). Thus, our present limited understanding of processes that determine potamoplankton composition currently prevents management agencies from developing predictive tools for
assessing ecological quality based on phytoplankton
community structure in large rivers. Attempts to use
planktonic algae as indicators of river water quality
using a community structure approach similar to
methods employed for benthic diatoms and ma-
PHYTOPLANKTON AND LARGE RIVER MANAGEMENT
croinvertebrates (Plafkin et al. 1989, Lowe and Pan
1996) have so far been unsuccessful (Williams 1964,
1972, del Giorgio et al. 1991). Studies have shown
that changes in phytoplankton composition reflect
not only variations in water quality, but also changes
in physical variables and biotic interactions. Variations in water chemistry may alter relative proportions of a few dominant taxa but often have little
effect on the overall assemblage.
The ability to apply phytoplankton-based assessments of the ecological status of large and lowland
rivers lies in a better understanding of the growth
rates and requirements of suspended algal species
in rivers. As some studies have shown (e.g. Kiss
1994), major changes in a watershed can be detected by biomass changes and by alteration of the
growth pattern of certain potamoplankton species.
Therefore, a system may be devised that evaluates
ecological change by contrasting the present community with reference data from unaltered stretches
or time periods. Moreover, as good historical data
are rarely available, the reference situation may be
simulated through a comprehensive river phytoplankton model, based on known individual river
phytoplankton growth submodels. The potamoplankton models currently available, however, are
still unable to simulate the potamoplankton composition under the dynamic conditions that prevail
in rivers, although the attempts made so far are
promising (Dauta 1983, Garnier et al. 1995, Cloot
and Le Roux 1997, Ruse and Love 1997). We suggest that the main difficulty in developing a potamoplankton model lies not in incorporating more
river–watershed features, examples of which already
exist (e.g. Smitz et al. 1997); rather, new models
must include information on physiological parameters of the algal taxa present. This has yet to be attempted comprehensively for any large river system
to date. While this goal may appear to be a very large
task, detailed lake phytoplankton models have been
developed using these principles quite successfully
(Hilton et al. 1992, Reynolds 1996, Reynolds and
Irish 1997). Therefore, river ecologists should build
on this knowledge and adapt new algal growth models for river environments.
Thanks go to Dr. James H. Thorp, Clarkson University, New York,
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