Download Trends in Estuarine Phytoplankton Ecology

7.02
Trends in Estuarine Phytoplankton Ecology
C Lancelot, Université Libre de Bruxelles, Brussels, Belgium
K Muylaert, K.U. Leuven Campus Kortrijk, Kortrijk, Belgium
© 2011 Elsevier Inc. All rights reserved.
7.02.1
7.02.2
7.02.2.1
7.02.2.2
7.02.2.3
7.02.2.4
7.02.2.5
7.02.3
7.02.4
References
Introduction
Multi-Stressors behind Phytoplankton Development in Estuaries
Hydrodynamics
Hydro-Sedimentary Processes as Drivers of Light Availability
Nutrients
Salinity Gradient
Top-Down Control
Trends
Conclusions and Perspectives
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Abstract
A review of the multi-stressors behind phytoplankton development in estuaries shows that differences in salinity tolerances
and in physical processes (freshwater and particle transport, tidal amplitude, mixing processes, and in upward/downward
boundaries) feature in the phytoplankton successions and magnitudes that are specific to each estuary. Nutrient loads play
generally only a small role except in subtropical and tropical estuaries during periods of low discharge. The high variability of
estuaries worldwide makes their monitoring difficult and the global assessment of their functioning in response to changes in
climate and anthropogenic pressures stresses the need to develop online fully coupled river–estuary–coastal ocean physical–
biological models at regional scales.
7.02.1 Introduction
Estuaries are shallow open systems strongly influenced by river
inflows, mixing with the coastal ocean and exchanges across
the sediment– and atmosphere–water interfaces. They have
distinct salinity gradients in their lower part and a tidal influ­
ence in the upper freshwater part providing specific
hydrological properties when compared to rivers. Estuaries
form transitional zones between freshwater and marine envi­
ronments and are characterized by a large variability in
physical, chemical, and biological properties under the dual
influence of climate and anthropogenic changes (e.g., Paerl
et al., 2010). Aquatic phototrophs (phytoplankton, periphy­
ton, and macrophytes) compete for light, nutrients and
inorganic carbon, and the balance among phototrophs changes
with size, depth, and nutrient status of their habitat. In the
often nutrient-rich and turbid shallow waters such as estuaries,
phytoplankton usually dominates over macrophytes and
benthic microalgae (Sand-Jensen and Borum, 1991).
The question of an actual estuarine phytoplankton assem­
blage has been debated lively. Recent statistical analysis of
ecological boundaries in estuaries shows no evidence of true
estuarine biological communities but rather the existence of a
continuum of biological assemblages along the salinity gradient,
defining estuaries as ecoclines rather than ecotones (Attrill and
Rundle, 2002; Quinlan and Philips, 2007; Muylaert et al., 2009).
While marine phytoplankton are adapted to high salinity and
freshwater species to low salinity, some of them have evolved at
an intermediate salinity (e.g., Jackson et al., 1987; Devassy and
Goes, 1988; Roubeix and Lancelot, 2008; Muylaert et al., 2009)
or are resistant to salinity fluctuations and thus can survive in
brackish waters (Figure 1). Understanding how freshwater and
marine phytoplankton organisms respond to the varying and
contrasted physico-chemical conditions encountered when
transported downstream and upstream along the estuary is a
prerequisite for describing the estuarine succession of phyto­
plankton assemblages and for assessing the ecological and
biogeochemical role of estuaries, in terms of, for example, reten­
tion of anthropogenic nutrients and direction of water–
atmosphere CO2 exchange.
Providing a connection with the hinterland through the
river, water quality and biological communities of estuaries
are strongly influenced by freshwater inputs that deliver sus­
pended sediments, biogenic elements (nutrients) like nitrogen
(N), phosphorus (P) and silicon (Si), and contaminants from
land runoff and wastewater discharge (see Chapter 5.10).
Nutrient inputs in particular reflect the ways that humans
produce and consume food and how they manage water­
courses and wastewaters in the watersheds. Nutrient inputs to
the tidal freshwater estuary and the response of the freshwater
phytoplankton community are, however, modulated by cli­
mate change acting on runoff and rainfall and also on
temperature and clouds. At the ocean boundary, changes in
atmospherically driven currents by modifying phytoplankton
composition and growth conditions in nearshore waters also
influence phytoplankton species dominance in the estuary
through the tidal flushing which transports coastal phytoplank­
ton cells into the estuary.
Phytoplankton development and species dominance in the
estuaries are clearly driven by a complex interplay between
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Trends in Estuarine Phytoplankton Ecology
Probability
1.0
(a)
0.8
Tetrastrum staurogeniaeforme
Scenedesmus spp.
0.6
Coelastrum sp.
Planktothrix aghardii
0.4
Gymnodinium sp.
Monoraphidium centortum
0.2
Pediastrum spp.
0
1.0
(b)
Stephanodiscus hantzschii
ocean–estuary–river physical–biogeochemical models that
describe nutrient/contaminant transformations and phyto­
plankton development along the land–ocean continuum
in response to regional changes in climate and human pres­
sures. As a first step in this direction, we review here the multistressors which are behind phytoplankton development in
estuaries in order to help to understand trends in estuarine
phytoplankton and to define the needed physical and trophic
resolution of the coupled models which are to be used for
assessing their future in response to changing climate and for
the implementation of environmental policies.
Probability
0.8
Rhodomonas sp. Cyclotella scaldensis
Thalassiosira proschkinae
0.6
Actinocyclus normannii
Melosira varians
Oocystis spp.
0.4
Melosira nummuloides
0.2
0
1.0
Probability
0.8
Actinastrum hantzschii
(c)
Guinardia delicatula
Raphoneis amphiceros
0.6
Lithodesmium undulatum
Chaetoceros
Rhizosolenia pungens
0.4
0.2
Mixing depth-to­
photic depth ratio
Triceratuim alternans Rhizosolenia sp.
0
16
14
12
10
8
6
4
2
0
30
(d)
(e)
Salinity
25
20
15
10
5
0
0
20
40
60
80
100 120 140 160 180
Distance from estuary mouth (km)
Figure 1 Modeled responses of selected phytoplankton taxa along the
longitudinal axis of the Scheldt estuary. Taxa with population maximum in
river (a), within the estuary (b), and in coastal waters (c) as adapted from
Muylaert et al. (2009) in relationship with the annual mean mixing-to-photic
depth (d) and salinity distribution (e) as adapted from Lionard, M.,
Muylaert, K., Van Gansbeke, D., Vyverman, W., 2005b. Influence of changes
in salinity and light intensity on growth of phytoplankton communities from
the Schelde river and estuary (Belgium/The Netherlands). Hydrobiologia 540,
105–115. With kind permission from Springer Science + Business Media.
processes occurring over the regional ocean basins and within
the watersheds. Such a complexity could be resolved by devel­
oping and implementing at regional scale coupled
7.02.2 Multi-Stressors behind Phytoplankton
Development in Estuaries
Considering the interplay between the stressors that are specific
for each estuary, each of them is discussed below as a necessary
but not exclusive condition for allowing phytoplankton to
develop in the estuary. When relevant, interacting factors are
pointed out and specific examples are given.
7.02.2.1
Hydrodynamics
Freshwater inputs in estuaries result in a net downstream trans­
port of water within the estuary. Together with freshwater
inputs, the exchange of water with the coastal zone determines
the residence time of the water within the estuary. Yet for
several of the world’s largest rivers, discharge often exceeds
the tidal prism and mixing of freshwater and seawater occurs
in coastal waters (e.g., Amazon (Demaster et al., 1996),
Changjiang (Edmond et al., 1985), Zaire/Congo (van
Bennekom et al., 1978), and Danube (Humborg et al.,
1997)). Conversely in the arid and subtropical regions, evapo­
ration exceeds precipitation during the dry season giving rise to
higher salinity in the estuary compared to the adjacent coastal
waters. This process typifies inverse estuaries such as the Fitzroy
estuary (Radke et al., 2010), the Spencer Gulf (Smith and Veeh,
1989), and the San Quintin Bay (Camacho-Ibar et al., 2003).
Because phytoplankton is passively transported along with
the water currents, it can only increase within the estuary when
net specific growth rates (i.e., the balance between phytoplank­
ton growth and losses by lysis, grazing, and/or sedimentation)
exceed the residence time of the water (Lucas et al., 2009). In
many estuaries, the residence time is primarily influenced by
river discharge (see Chapter 5.11). Hence, the development of
phytoplankton blooms is often, but not always, inversely cor­
related with river discharge (e.g., Strayer et al., 2008). River
discharge varies strongly over different timescales and latitudes
and this variability affects phytoplankton dynamics. As a gen­
eral trend, the timing and magnitude of hydrological factors is
more contrasted, episodic, and intense in subtropical and trop­
ical estuaries due to the seasonal monsoons and the occurrence
of droughts, tropical storms, and hurricanes (e.g., Devassy and
Goes, 1988; Eyre and Balls, 1999; Briceño and Boyer, 2010).
Discharge differences between dry and wet seasons are gener­
ally two orders of magnitude higher than those observed in
temperate estuaries (Eyre and Balls, 1999).
In temperate regions, river discharge tends to be higher
during winter and lower in summer. In such estuaries, the
onset of the phytoplankton bloom in spring often coincides
Trends in Estuarine Phytoplankton Ecology
with a decrease in river discharge. In the San Francisco Bay
estuary, phytoplankton blooms are restricted to periods of
low river discharge in summer (Cloern et al., 1983). In the
turbid freshwater zone of macrotidal estuaries, phytoplankton
blooms are also restricted to sustained periods of low river
discharge in summer (Filardo and Dunstan, 1985; Jackson
et al., 1987; Sin et al., 1999; Muylaert et al., 2005; Arndt
et al., 2007). River discharge is sensitive to meteorological
conditions showing contrasted inter-annual variability that is
often reflected in phytoplankton biomass. In the lower reaches
of the Hudson River estuary, phytoplankton blooms are
limited by a short residence time of the water. In years when
discharge of the Hudson River is low however, residence time
in the estuary increases and phytoplankton blooms can
develop (Howarth et al., 2000). In the turbid freshwater zone
of temperate macrotidal estuaries, phytoplankton biomass also
tends to be higher during dry than wet summers (e.g., Lionard
et al., 2008).
In tropical estuaries the monsoon-driven seasonal discharge
patterns and the episodic pulses of freshwater inputs caused by
tropical storms or hurricanes constrain phytoplankton biomass
building for short periods immediately following floods, that
is, when the turbidity caused by the transported suspended
particles is decreased (e.g., Devassy and Goes, 1988; Sarma
et al., 2009). The occurrence of drought/flood events strongly
influences the bloom pattern, that is, the position and species
dominance of the phytoplankton biomass maxima along the
estuary (e.g., Valdes-Weaver et al., 2006; Costa et al., 2009).
Such a positive correlation between phytoplankton increase
and transport time is still not found in some estuaries (see
references in Lucas et al., 2009). This is explained by a negative
balance of phytoplankton growth versus loss processes due to
either elevated grazing (Alpine and Cloern, 1992; Strayer et al.,
2008), lysis, or sedimentation rates and suggests that the phy­
toplankton growth-loss balance is modulating the relationship
between phytoplankton development and hydrology (Cloern,
1996). Hence, the different response of phytoplankton taxa
such as cyanobacteria and diatoms to changing transport
times reported by Paerl et al. (2006) might be explained by
differences in species growth physiology and susceptibilities to
grazing and sinking (Lucas et al., 2009).
7.02.2.2 Hydro-Sedimentary Processes as Drivers of Light
Availability
Light is a necessary condition for phytoplankton growth. Light
availability in the water column is determined by incident sur­
face light and the vertical light extinction coefficient KPAR which
determines the depth of the euphotic zone. The ratio between
the euphotic depth and the upper mixing depth then deter­
mines the light available to phytoplankton cells when
transported up and down in the mixed layer. When the
euphotic depth is shallow compared to the mixing depth,
photosynthesis is insufficient to compensate for dark respira­
tion and phytoplankton cells lyse or enter a dormancy phase.
Several studies have illustrated the importance of evaluating
the mixing-to-euphotic depth ratio for understanding light
limitation of phytoplankton development in estuaries
(Kromkamp and Peene, 1995; Irigoien and Castel, 1997;
Desmit et al., 2005). In estuaries, KPAR is mainly determined
by the elevated turbidity that distinguishes many estuaries from
7
freshwater and ocean ecosystems where light attenuation is
primarily governed by phytoplankton cells. Turbidity has long
been considered as the main factor controlling light availability
and hence phytoplankton development in estuaries, irrespec­
tive of nutrient concentrations (Cloern, 1987; Fichez et al.,
1992; Heip et al., 1995). This elevated turbidity is due to high
concentrations of suspended sediments driven by essentially
three mechanisms acting at different locations along the estu­
ary: (1) tidal currents or wind or wave action resulting in
resuspension of bottom sediments in shallow estuaries;
(2) local accumulation of suspended matter at the so-called
turbidity maxima; and (3) river inputs of large quantities of
suspended sediments of terrestrial origin into the estuary.
Tidal currents are particularly strong in estuaries that are
situated relatively far from amphidromic points such as those
near the English Channel or along the East coast of Canada. By
causing resuspension of bottom sediments, the tidal currents
have a strong influence on turbidity and hence on light avail­
able for phytoplankton photosynthesis. Monbet (1992)
showed that turbidity is higher and phytoplankton biomass is
significantly lower in macrotidal estuaries (>2 m tidal range)
compared to microtidal estuaries (<2 m tidal range). Examples
of macrotidal estuaries where phytoplankton growth is pre­
dominantly light-limited by high turbidity are the San
Francisco Bay estuary (Cloern, 1987), the tropical Fitzroy estu­
ary (Radke et al., 2010), the Schelde estuary (Soetaert et al.,
1994), and the Gironde estuary (Irigoien and Castel, 1997). In
macrotidal estuaries, the tidal range may vary substantially over
bi-weekly timescales following the spring–neap tidal cycle. This
spring–neap tidal variability is reflected in turbidity and influ­
ences phytoplankton growth. In the San Francisco Bay estuary,
the onset of phytoplankton blooms often coincides with neap
tides while blooms are often terminated during spring tides
(Cloern, 1991).
The typical funnel shape of many estuaries often induces a
magnification of the tidal wave, resulting in higher tidal amplitude
in the upper freshwater estuary than in the nearby coastal zone.
Extreme examples of this phenomenon are the Bay of Fundy in
Eastern Canada and the Bristol Channel in England, where the
tidal range exceeds 15 m. As a result, turbidity and hence KPAR are
often higher in the upper tidal freshwater (e.g., Chubut river
estuary: KPAR = 4–6 m−1; euphotic depth = 1.2–0.8 m; Helbling
et al., 2010) than the marine (KPAR = 1–2 m−1) region. In the tidal
freshwater estuary, phytoplankton growth is only possible in
shallow zones (e.g., the Hudson River estuary (Cole et al.,
1992); the Scheldt freshwater estuary (Desmit et al., 2005;
Muylaert et al., 2005)). Nevertheless, during summer at low
discharge and longer residence time, dense phytoplankton
blooms may develop in these turbid systems, with chlorophyll
a concentrations exceeding 100 mg m−3, that is, much higher
than in more downstream zones (Filardo and Dunstan, 1985;
Jackson et al., 1987; Schuchardt and Schirmer, 1991; Muylaert
et al., 2005).
In estuaries that are situated close to amphidromic points,
tidal currents are too weak to resuspend bottom sediments. In
these estuaries, turbidity is mainly caused by wave and/or wind
resuspension. As an example, wind-induced turbidity has been
shown responsible for large day-to-day variability in phyto­
plankton production in San Antonio Bay (Macintyre and
Cullen, 1996).
8
Trends in Estuarine Phytoplankton Ecology
Maximum turbidity zones are common in relatively long
macrotidal estuaries (Uncles et al., 2002) and are formed by
two mechanisms: (1) particle accumulation in the residual
circulation near the salt wedge of partially stratified estuaries
and (2) tidal pumping of particles in estuaries with an asym­
metrical tidal cycle. Maximum turbidity zones in estuaries are
often in regions where primary production is strongly lightlimited (e.g., Soetaert et al., 1994; Radke et al., 2010). In
addition to salinity stress, light limitation in the turbidity max­
imum may constitute a barrier for upward (or downward)
transport of marine (or freshwater) phytoplankton taxa.
Despite the severe light limitation, increased phytoplankton
biomass has been reported in the maximum turbidity zone of
some estuaries (e.g., San Francisco Bay estuary (Cloern et al.,
1985); the Ouse River estuary (Uncles et al., 1998); and the
Fitzroy estuary (Radke et al., 2010)). Most likely, this is caused
by physical accumulation rather than physiological growth.
Physical accumulation of phytoplankton cells in the maximum
turbidity zone may be the result of the same mechanisms by
which sediment particles accumulate, that is, by trapping in the
residual circulation near the salt wedge or by tidal pumping.
Some species from the diatom genus Thalassiosira that tend to
be associated with the maximum turbidity zone are often cov­
ered with a layer of sediment particles, possibly because this
enhances trapping of the cells in the turbidity maximum and
avoids export of the cells out of the estuary (Cloern et al., 1983;
Muylaert and Sabbe, 1996).
In turbid estuaries, phytoplankton often spends most of the
time in darkness and photosynthesis only occurs in short inter­
mittent periods when phytoplankton is transported in the
shallow euphotic zone. In the absence of grazers, phytoplank­
ton growth is only possible if respiration rates in the darkness
are assumed to be very low (e.g., Kromkamp and Peene, 1995).
Several studies have indeed shown that phytoplankton cells
can survive prolonged time in total darkness, even up to several
months (Dehning and Tilzer, 1989; Wasmund, 1989), which
implies a potential for severe reduction of dark respiration
rates. It is not always clear whether phytoplankton cells survive
dark periods in a state of dormancy or by heterotrophy. It has
long been known that many phytoplankton species are capable
of taking up simple organic substrates, a heterotrophic nutri­
tion mode that is referred to as osmotrophy (Lewin and
Hellebust, 1970; Lylis and Trainor, 1973; Berman et al., 1977;
Hellebust, 1978). When present at sufficiently high concentra­
tions, such organic molecules can support phytoplankton
growth at rates equal to or even higher than in the light
(Richmond, 2004). In natural ecosystems, however, the con­
centration of simple organic molecules is too low
(see Chapters 5.02 and 5.03) to sustain osmotrophic growth
of phytoplankton. Moreover, the affinity of phytoplankton for
such substrates is too low to compete with bacteria (Darley
et al., 1979) which are otherwise abundant in estuaries and
especially in turbid waters. Supporting this, recent mesocosm
experiments demonstrated glucose assimilation by Scheldt
estuary phytoplankton but this uptake was much lower than
the organotrophic activity associated with ambient bacteria
(Van den Meerssche, 2009). Interestingly enough recent studies
have shown phytoplankton cells metabolizing complex or
recalcitrant organic molecules (Tuchman et al., 2006;
Blackburn et al., 2009). In particular, the freshwater green
algae Chlamydomonas has recently been shown to metabolize
aromatic photo-degradation products of terrestrial origin
(Tittel et al., 2009). The generally high inputs of continental
organic matter into estuaries and the possible photo-degrada­
tion at the surface of the estuarine continuum may provide
phytoplankton cells with a continuous supply of suitable
energy substrates, supporting their energy demands when
exposed to prolonged periods of darkness.
Turbidity in estuaries not only regulates phytoplankton
growth but also controls the community composition, with
diatoms better adapted to the turbid conditions that are typical
of many estuaries, possible mixotrophs in maximum turbidity
zones, and pico-/nano-phytoplankton in low turbid oligotro­
phic waters. In turbid estuaries, diatoms tend to dominate the
phytoplankton community throughout the year when
nutrients are sufficient (e.g., Schuchardt and Schirmer, 1991;
Muylaert et al., 2000). In estuaries where turbidity is low,
diatoms tend to be dominant in spring, when light levels are
relatively low and stratification does not occur (Marshall et al.,
2006) but dinoflagellates and cyanobacteria are usually more
abundant in summer during periods of low discharge, long
residence time and minimal flushing rates and turbidity
(Valdes-Weaver et al., 2006).
7.02.2.3
Nutrients
Besides sufficient light, nutrients are the second necessary con­
dition for the biosynthesis of new phytoplankton cells. The
potential magnitude of phytoplankton bloom is set by the sup­
ply rate of the least abundant nutrient either nitrogen (N) or
phosphorus (P) or silicon (Si) with the latter for the only diatom
growth. Many estuaries are nutrient rich because of new inputs
from the land surface and remineralization processes occurring
in the maximum turbidity zone and sediments. The potential
primary production is not however always reached either
because estuaries are very turbid imposing light limitation on
algal production or because the water renewal rate of the system
is very fast. Therefore, phytoplankton’s response to nutrient
loads differs from estuary to estuary, from segment to segment,
and from time to time within any segment of an estuary.
Despite this, an analysis of 40 temperate and subtropical
estuaries (Monbet, 1992) shows a clear distinction between
microtidal and macrotidal estuaries with the former being much
more sensitive to nutrient loads that allow a higher maximum
biomass to be reached. The strong tidal influence on water col­
umn mixing and transport and accumulation of suspended
sediments prevailing in macrotidal estuaries limits light availabil­
ity such that only a fraction of the nutrients imported by the river
are used by phytoplankton for their growth. The exported
nutrients usually sustain bloom development in the receiving
adjacent coastal waters, where turbidity is lower and residence
time of the water is longer. The best examples are to be found in
European coastal waters bordering the English Channel.
In estuaries with long retention times and a favorable
underwater light climate for photosynthesis, phytoplankton
can use most of the available nutrients. In such systems,
nutrient limitation is commonly observed and phytoplankton
blooms rely on seasonal and episodic nutrient inputs (Paerl
et al., 2010). Nutrient inputs from rivers vary seasonally and
inter-annually and this variability is often reflected in phyto­
plankton biomass. In Chesapeake Bay, a high river flow in
spring is often followed by high phytoplankton biomass in
Trends in Estuarine Phytoplankton Ecology
summer (Jordan et al., 1991). In the North Carolina estuarine
system, wet winters result in high nitrogen inputs into the
estuary, resulting in higher than average phytoplankton bio­
mass in the lower estuary (Mallin et al., 1993). On the contrary
during periods of low discharge, phytoplankton is often
limited by low nutrient concentrations. For instance, in the
North Carolina estuarine system (which includes the Neuse
River estuary and Pamlico Sound), phytoplankton tends to be
N-limited, with P occasionally being co-limiting during periods
of low river discharge (Mallin, 1994). In these systems, nutrient
limitation can be relieved by episodic pulsed inputs of
nutrients from runoff produced in the local watershed during
storms (Mallin et al., 1993; Paerl et al., 2001), direct inputs of
atmospheric N from rainfall (Paerl et al., 1990) or the transport
of nutrients from the coastal ocean during upwelling events
(e.g. Lara-Lara et al., 1980).
7.02.2.4
Salinity Gradient
The salinity gradient resulting from the mixing of freshwater
and seawater is a unique hydrological aspect of estuaries.
Salinity is often reported as the major factor determining phy­
toplankton distribution in estuaries, most species being
adapted to high, low, or intermediate salinities, and few species
occurring along the entire salinity range. Picoplankton, which
has been reported in high concentration in transitional waters
and survives in salinity between 5 and 87, appears to be an
exception (Paoli et al., 2007).
During downstream transport in the estuary, freshwater
species experience drastic salinity change. Stenohaline river
species (e.g., Asterionella formosa) lyse early in the salinity gra­
dient because their specific limit of salinity tolerance is
exceeded, as evidenced by the high phytoplankton mortality
reported in the oligohaline region (salinity <5) of estuaries
(Morris et al., 1978; Ahel et al., 1996; Ragueneau et al.,
2002). On the contrary, some euryhaline species are able to
regulate their intra-cellular osmotic pressure and grow at a
higher salinity further seaward in estuaries (Flameling
and Kromkamp, 1994; Orive et al., 1998). Among these holo­
euryhaline species, the most common is the cosmopolite dia­
tom Cyclotella meneghiniana which is able to grow from
freshwater to marine conditions (Carpelan, 1978) reaching an
optimum around salinity 18 (Tuchman et al., 1984; Roubeix
and Lancelot, 2008). Similarly, marine species that are
advected upstream in the estuary by tidal mixing disappear
when salinity decreases down to 10 (e.g., Roubeix et al., 2008).
Although the disappearance of marine and freshwater spe­
cies at intermediate salinities is often ascribed only to osmotic
stress, recent experiments with pure phytoplankton cultures as
well as with natural estuarine phytoplankton communities
suggest that light availability may also contribute to the tran­
sition from a freshwater to a marine community along the
salinity gradient (Flameling and Kromkamp, 1994; Lionard
et al., 2005b). As the salinity gradient is often also the location
of turbidity maximum, the disappearance of marine and fresh­
water phytoplankton cells near the salinity gradient (Muylaert
and Sabbe, 1999) may also be impaired by light limitation.
Model reconstruction of an estuarine salinity gradient, making
use of three diatom species (stenohaline freshwater (SR),
euryhaline freshwater (ER) and euryhaline marine (EM)), shows
indeed that along with freshwater and marine species initial
9
conditions, residence time, light availability (Figure 2(b)), and
salinity tolerance (Figure 2(c)) are all important in determining
the species succession and the magnitude of their bloom along
the estuary (Roubeix et al., 2008).
7.02.2.5
Top-Down Control
Top-down control of estuarine phytoplankton includes grazing
by pelagic (meso-zooplankton and micro-zooplankton) and
benthic herbivores. The meso-zooplankton in estuaries is com­
posed of micro-crustaceans, of which the dominant groups are
calanoid copepods in marine and brackish waters and cyclo­
poid copepods and cladocerans in freshwater. Due to the
relatively long generation time of meso-zooplankton, their
biomass in estuaries is limited by the low residence time of
the water (Pace et al., 1992). Yet copepods have been reported
to escape washing out of the estuary by migrating vertically in
the water column to avoid downstream-directed ebb currents
(Kimmerer et al., 1998). Meso-zooplankton is sensitive to
anoxic conditions and may be absent in extremely polluted
estuaries with low-oxygen concentrations (Appeltans et al.,
2003). Although phytoplankton is important as a food source
for meso-zooplankton in estuaries, including the turbid
ones (e.g., Tackx et al., 2003), the grazing pressure of meso­
zooplankton on phytoplankton in estuaries is generally low
compared to other ecosystems (White and Roman, 1992).
The micro-zooplankton comprises nauplius stages of
copepods, and rotifers and protozoa (ciliates and heterotrophic
dinoflagellates). Because their growth rates are similar to those
of phytoplankton, the grazing pressure of micro-zooplankton
is expected to be significant. This is the case in the lower reaches
of some estuaries such as the lower Hudson River (Lonsdale
et al., 1996) and the San Francisco Bay and the Chesapeake Bay
(Murrell and Hollibaugh, 1998) but not in freshwater tidal
estuaries such as those of the Potomac tributary of
Chesapeake Bay and the Schelde estuary (Lionard et al.,
2005a; Sellner et al., 1993), probably because micro-sized
diatoms or filamentous cyanobacteria dominate the freshwater
assemblages.
Benthic filter-feeders may, in some estuaries, be important
grazers of phytoplankton (Herman 1993; Lucas et al., 1998). In
Danish estuaries, biomass of benthic bivalves (mainly Mytilus
edulis) is negatively correlated with chlorophyll a concentra­
tions in the water column (Conley et al., 2000). A numerical
study exploring the phytoplankton dynamics in a shallow
tidally mixed estuary in the presence of benthic grazers points
out that tidal variability by determining the height of the water
column can represent maximum phytoplankton production or
consumption, depending on the light availability/grazing bal­
ance (Lucas et al., 1999). Supporting this, it has been suggested
that increases in phytoplankton biomass during water-column
stratification are partly due to the fact that phytoplankton is
released from grazing by benthic bivalves (Møhlenberg, 1995;
Carstensen et al., 2007). The control by benthic bivalves on
phytoplankton biomass is particularly clear when the impact of
accidentally introduced bivalves is evaluated. Within 2 years,
the introduction of the clam Potamocorbula amurensis in the San
Francisco Bay estuary in 1986 reduced phytoplankton develop­
ment by a factor of 2 (Alpine and Cloern, 1992). Dreissena
polymorpha, first observed in 1991 in the freshwater zone of
the Hudson River estuary, was shown to be filtering the entire
25
100
20
80
15
60
ER
SR
10
40
5
20
EM
0
0
Diatoms (µmol-Si l–1)
0
5
10
15
20
25
30
25
100
20
80
15
60
10
40
5
20
0
0
0
5
10
15
20
25
30
(c) 25
Diatoms (µmol-Si l–1)
Nutrient (µmol-Si l–1)
Diatoms (µmol-Si l–1)
(a)
Nutrient (µmol-Si l–1)
Trends in Estuarine Phytoplankton Ecology
100
20
80
15
60
10
40
5
20
0
Nutrient (µmol-Si l–1)
10
0
0
5
10
15
20
25
30
Salinity
Figure 2 Model simulation showing the evolution of stenohaline freshwater (SR) and euryhaline freshwater (ER) and marine (EM) diatoms in salinity
gradient for (a) sufficient light, (b) limiting light, and (c) in the absence of ER. The gray line shows the concomitant evolution of silicic acid DSi. Adapted
from Roubeix, et al., 2008.
estuary volume daily in 1992 (Strayer et al., 1996) resulting in a
85% drop in phytoplankton biomass (Caraco et al., 1997).
Altogether, the introduction of bivalves in the San Francisco
Bay and the Hudson River also had a negative effect on meso­
zooplankton (Caraco et al., 1997).
7.02.3 Trends
Human activity is directly and indirectly affecting estuarine
processes, which has implications for the functioning of the
adjacent coastal waters. This, in turn, is influencing biological
processes in the lower estuary through the upward advection of
active coastal marine species (Cloern and Jassby, 2008).
Human pressures in the watershed (population density, land
use, agricultural practices, industrial activities, wastewater treat­
ment, river damming/rectification, and irrigation) and the
biogeochemical functioning of the river system determine the
level of nutrients (N, P, and Si), suspended particles, and fresh­
water species entering the estuary (e.g., Billen and Garnier,
1999). Between 1960 and 1990, the input of N and P (relative
to Si) to estuaries and coastal seas of wealthier countries
increased tremendously in response to increased population
density, and economical development (e.g., Billen et al., 2001,
2005; Boesch, 2002) leading to undesirable effects like harmful
algal blooms and related oxygen depletion or foam events
(e.g., Lancelot, 1995; Pearl et al., 2006). This phenomenon,
designated as cultural eutrophication, is now observed in
Trends in Estuarine Phytoplankton Ecology
developing countries (e.g., major Chinese estuaries including the
Changjiang and Zhujiang; Zhang, 1996; Liu et al., 2009) while
nutrient inputs to European (e.g., Soetaert et al., 2006) and
North-American (e.g., Paerl et al., 2006) estuaries and coastal
seas have decreased, especially P because of its banning in wash­
ing powders and the implementation of wastewater treatment
plants with so far, however, little visible effect on eutrophication
symptoms.
These nutrient modifications occurred in a period of chang­
ing climate as well. The ways that changing nutrients are
processed by phytoplankton differ between stratified and tur­
bid estuaries and are modulated by climate variability. Climate
constrains cloud cover, temperature, river discharge, sediment
transport, and deposition/resuspension processes that together
determine the available light to the transported cells.
Estuaries where phytoplankton production is strongly
limited by light or by residence time rarely display signs of
eutrophication. In such systems, only a fraction of the nutrients
imported by the river are used by phytoplankton and eutrophi­
cation manifests in the coastal zone, where turbidity is lower
and residence time of the water is longer. Examples are to be
found in North European macrotidal estuaries such as the
Scheldt or the Seine where undesirable eutrophication is
observed in spring in the receiving coastal waters
(e.g., Lancelot et al., 2007).
In stratified estuaries and in estuaries with long residence
times and a favorable underwater light climate phytoplankton
can use most of the available nutrients (Figure 2). In such
systems, nutrient limitation of phytoplankton biomass is com­
monly observed. This has been particularly documented for
large North American estuaries such as the Chesapeake Bay,
the Delaware, and the North Carolina estuarine systems
(Rudek et al., 1991; Fisher et al., 1992; Mallin, 1994; Pennock
and Sharp, 1994; Malone et al., 1996). Those low-turbid estua­
ries where phytoplankton development is limited by nutrient
availability are sensitive to quantitative and qualitative changes
in nutrient loads.
The increased load of N and P, while Si is maintained, is
often associated with a shift in the phytoplankton community
dominance toward nuisance or harmful nonsiliceous species
such as Pfiesteria piscicida in the Chesapeake Bay (Burkholder
et al., 1992), Alexandrium tamarense in the St. Lawrence estuary
(Fauchot et al., 2005), mixed dinoflagellate blooms in North
Carolina estuaries (Mallin, 1994), and Prorocentrum minimum in
Danish estuaries (Carstensen et al., 2007). Similarly, cyanobac­
teria blooms are reported in the freshwater zones as Microcystis
blooms in the Potomac River tributary of Chesapeake Bay
(Sellner et al., 1993), the Guadiana estuary (Rocha et al.,
2002), and the North Carolina estuarine system (Christian
et al., 1986). In turbid estuaries, harmful algal blooms are
often absent despite often very high nutrient concentrations.
Future changes in riverine nutrient delivery due to improved
environmental management are therefore expected to affect
phytoplankton development in estuaries where available
light and the water residence time are sufficient for allowing a
net growth. This is where future regional climate change
might have an influence by modulating the river discharge
(i.e., droughts vs. floods), the tidal amplitude, and storm
events. For instance, the drier and warmer summers expected
for North Europe might decrease river discharge and sediment
transport in the upper estuaries and improve conditions for
11
freshwater phytoplankton blooming with increased contribu­
tion of harmful events and nutrient depletions. Most freshwater
phytoplankton will, however, lyse when salinity increases and
the remineralized nutrients from dead cells will sustain growth
of marine phytoplankton in the lower estuary. On the contrary,
the expected wetter winter/spring will prevent growth in the
estuary and nutrients will be processed in the coastal zone
when favorable light conditions are reached. Climate variabil­
ity will also determine the composition of the biological
assemblage and growth conditions at the ocean boundary,
which will have an impact on the functioning of the lower
estuary. Such ocean–estuary connectivity was recently reported
for the San Francisco Bay where a change in the California
Current System driven by a cold phase of the East Pacific was
triggering phytoplankton blooms (Cloern et al., 2007). This
unexpected bloom event, which occurred during a period of
nutrient load reduction, was due to a trophic cascade initiated
by the intrusion of top predators of benthic bivalves, thus
suppressing their pressure on phytoplankton. Expected
increased temperature and changing wind stress will affect the
coastal ocean hydrodynamics with possible effects on phyto­
plankton species distribution and growth conditions (light,
temperature, and nutrients) that will impact on the functioning
of macrotidal estuaries. As suggested by Cloern et al. (2007),
estuaries have to be viewed as real interfaces being under the
influence of processes occurring in both the river drainage
basin and the coastal ocean, both of which are under the
control of climate variability while human pressures are mainly
located on the watershed.
7.02.4 Conclusions and Perspectives
Estuaries are all different but similar in their functioning.
Similarities in estuarine phytoplankton successions are the
result of the progressive replacement of freshwater species by
marine species when salinity is increased. However, their
respective contributions and positions along the salinity gra­
dient differ between estuaries. Both freshwater and marine
species obey physiological and ecological principles, their
growth being regulated by temperature, light availability,
nutrients, and grazing. Both freshwater and marine assemb­
lages possibly include euryhaline species that resist, to some
extent, the osmotic stress in the estuarine mixing zone, permit­
ting bloom occurrence at intermediate salinities when light is
sufficient. Little is known about the contribution of euryhaline
species to freshwater and marine assemblages but deserves
further attention considering the results obtained by a sensitiv­
ity analysis on their initial conditions performed with an
idealized stratified estuarine model (Roubeix et al., 2008;
Figures 2(a)–2(c)). In addition, mixotrophy as a survival strat­
egy in turbid estuaries needs to be explored.
As a general trend, the physical environment, more specifi­
cally the underwater light climate, and the water residence time
control phytoplankton development in estuaries. Typically,
conditions that favor phytoplankton blooms are low mixing­
to-photic depth ratios and low water discharges relative to the
estuarine volumes (Figure 3). The former may result from a
decrease in mixing depth that relies on the estuary morphology
but is influenced by stratification, and/or from a decrease in
turbidity. Mechanisms controlling turbidity are complex,
12
Underwater light availablity
Euphotic depth
Mixing depth
Trends in Estuarine Phytoplankton Ecology
Water retention time
River discharge
Estuarine volume
Figure 3 Scheme illustrating the control of phytoplankton biomass by
water retention time and the underwater light climate in estuaries. Water
retention time increases with estuarine volume and decreases with river
discharge. The underwater light availability increases with increasing
euphotic depth and decreases with increasing mixing depth. Dark area
corresponds to conditions favoring high phytoplankton biomass.
including import and flocculation of terrestrial sediments, tur­
bulence created by tidal currents or wind/waves and by mixing
of freshwater and seawater. The water residence time is influ­
enced by the morphology of the estuary and is controlled by
river discharge and seawater inflow. While the morphology of
the estuary is relatively fixed, river discharge and marine inflow
are highly variable and fluctuate at daily, seasonal, and interannual timescales. Because of this strong physical control of
phytoplankton growth, elevated nutrient concentrations in
estuaries often do not result in visible signs of eutrophication.
Signs of eutrophication are only apparent in estuaries where
primary production is not severely constrained by light climate
and/or water residence time.
Differences between estuaries are caused by those physical
properties that determine the freshwater and particles trans­
port, the tidal amplitude and the mixing processes, as well the
upward (river) and downward (coastal ocean) boundaries.
Worldwide, estuaries are constrained by different prevailing
climate, human pressures, and their interactions, making each
estuary specific in terms of water residence time, nutrient level,
turbidity (light availability), and freshwater/marine species
succession. As an example Figures 2(a) and 2(b) compare the
effect of light conditions on the development of SR and ER and
EM diatoms simulated in an estuarine gradient. While the SR
diatoms are little affected in his magnitude and disappear at
salinity 6, light limitation delays to higher salinities the devel­
opment of ER and EM that are more competitive for the uptake
of nutrient when light is sufficient (Figures 2(a)–2(c)). The
presence (or absence) of ER species has a strong impact on
the distribution and magnitude of freshwater and marine spe­
cies (Figures 2(a)–2(c)). The absence of ER species allows the
marine species penetrating the estuary to bloom and deplete
nutrients.
As the interface between the river and the ocean, estuaries
provide complex and fluctuating habitats for freshwater and
marine phytoplankton. This high variability complicates the
monitoring and assessment of the functioning of estuaries in
response to changing climate and anthropogenic pressures.
Little is known about, for example, estuary responses when
physical changes (e.g., increase/decrease in water residence
time and turbidity), climate (e.g., modification of freshwater
inflow, temperature, and clouds), or top-down effects
(removal/import of filter-feeders) are superimposed onto
changing nutrient emissions. Due to the high interest in coun­
teracting cultural eutrophication, the estuary has long been
considered as part of the unidirectional land–ocean contin­
uum. As suggested by Cloern et al. (2007), one cannot ignore
the effect of climate change on the functioning of the coastal
ocean and the subsequent impact on the lower estuary. It is
therefore urgent to set up online-coupled river–estuary–ocean
physical–biogeochemical models of high resolution for
describing the effect of changing human pressure in a changing
climate. Such models exist for the river drainage basin, the
estuary, and the adjacent coastal sea, and have been partially
coupled (e.g., Cugier et al., 2005; Lancelot et al., 2007; Arndt
et al., 2011) but were never coupled online, that is, allowing
freshwater and marine plankton species to develop when trans­
ported downward and upward in the estuary. This constitutes a
challenge for the next decade that will need new numerical
developments supported by new eco-physiological studies for
model equations’ parametrization and field observations for
their validation.
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