Chapter 11 Dynamics of Seagrasses
289
of seagrass globally (Short and Wyllie-Echeverria,
1996). The newly published World Atlas of Seagrasses provides a global synthesis of the distribution
and present status of seagrass meadows and documents that seagrasses are being steadily destroyed by
the run-off of nutrients and sediments from land and
by boating, land reclamation, dredging, and some
fishing methods (Green and Short, 2003). Along
with increased eutrophication, negative cascading
effects upon the loss of seagrass biomass are common. These involve, for example, increased resuspension of sediments and thereby increased turbidity that further reduces seagrass abundance (Duarte,
1995). Moreover, the occurrence of anoxia during
warm calm periods becomes more frequent as eutrophication increases (Rabalais and Turner, 2001)
and may seriously affect seagrasses (Terrados et al.,
1999; Greve et al., 2003) and cause diebacks (Rask
et al., 2000; Plus et al., 2003).
One example of seagrass decline upon increased
eutrophication is from the Dutch Wadden Sea. Both
the fact that littoral eelgrass gradually disappeared
after the mid-1960s and the fact that sublittoral eelgrass beds failed to recover from the wasting disease
have been interpreted as responses to increased turbidity caused by eutrophication (Giesen et al., 1990).
Florida Bay also experienced a serious loss of seagrasses over a decade (1984–1994), which was partly
due to increased turbidity (Hall et al., 1999) and in
Chesapeake Bay losses or Zostera marina and Ruppia maritima were also related to increased turbidity as a result of eutrophication (Orth and Moore,
1983). In Waquoit Bay, Massachusetts Short and
Burdick (1996) related housing development and nitrogen loading to eelgrass habitat loss over the period
1987–1992 (Fig. 10). The effect occurred largely via
ground water and resulted in a gradual fragmentation
and loss of the meadows.
Examples of recolonization upon reduction of eutrophication are limited. The seagrass cover in Cockburn Sound, Western Australia, was markedly reduced between 1976 and 1981 as a response to
eutrophication, but reductions in nutrient loads in
the 1980s did not lead to recolonization (Walker
et al., Chapter 23). It is likely that alterations in
shelf-environments during the period without seagrasses have rendered the area unsuitable for seagrass growth (Kendrick et al., 2002). In contrast,
Posidonia coriaceae and Amphibolis griffithii have
recolonized former seagrass areas in Success Bank,
Western Australia, at surprizingly high rates in100
75
50
25
0
0
5000
10000
15000
20000 25000
Nitrogen loading (kg km -2 yr -1 )
Eelgrass area (%)
Log(y) = 1.648 - 0.000044x
R 2 = 0.888
1987
1988
1989
Fig. 10. Comparison of nitrogen loading rates and eelgrass area
in the Waquoit Bay estuary’s sub-basins over the first three years
of study. The log of eelgrass area is regressed against loading.
From Short and Burdick (1996), with permission from the Estuarine Research Federation.
volving both vegetative and sexual reproduction
(Kendrick et al., 1999; Walker et al., Chapter 23). An
extremely rapid eelgrass recolonization was also observed in the Archipelago of Southern Funen, Denmark. This area experienced an 80% reduction in
the distribution area of eelgrass following an anoxic
event during a warm summer period, but recovered
completely within 3 years through a combination of
vegetative growth of surviving shoots and germination of seeds (Rask et al., 2000). An even faster recolonization of Z. marina after anoxia-induced mortality was observed in the Thau Lagoon, French Mediterannean Sea (Plus et al., 2003).
Rapid recolonization seems possible if the disturbance causing the seagrass decline is limited in time
and space and if seedlings originating from the sediment bank or from neighbouring populations experience suitable growth conditions the following
year. By contrast, recovery of seagrass populations
from catastrophic decline on the landscape scale requires patch initiation from seeds transported from
distant populations and subsequent patch growth.
The survival chances of these initial patch stages
are low, and the formation of new extended patches
may, therefore, be a protracted process. Simulation
models show that small species with large recolonization potentials may recover within a few years
after a disturbance, while large species with small
recolonization potentials may require centuries to recover if the process is at all reversible (Duarte, 1995).
Colonization may be further delayed or impeded by
negative cascading effects (Duarte, 1995).
In many cases, declines of seagrass meadows are
not detected before marked losses have occurred
289
of seagrass globally (Short and Wyllie-Echeverria,
1996). The newly published World Atlas of Seagrasses provides a global synthesis of the distribution
and present status of seagrass meadows and documents that seagrasses are being steadily destroyed by
the run-off of nutrients and sediments from land and
by boating, land reclamation, dredging, and some
fishing methods (Green and Short, 2003). Along
with increased eutrophication, negative cascading
effects upon the loss of seagrass biomass are common. These involve, for example, increased resuspension of sediments and thereby increased turbidity that further reduces seagrass abundance (Duarte,
1995). Moreover, the occurrence of anoxia during
warm calm periods becomes more frequent as eutrophication increases (Rabalais and Turner, 2001)
and may seriously affect seagrasses (Terrados et al.,
1999; Greve et al., 2003) and cause diebacks (Rask
et al., 2000; Plus et al., 2003).
One example of seagrass decline upon increased
eutrophication is from the Dutch Wadden Sea. Both
the fact that littoral eelgrass gradually disappeared
after the mid-1960s and the fact that sublittoral eelgrass beds failed to recover from the wasting disease
have been interpreted as responses to increased turbidity caused by eutrophication (Giesen et al., 1990).
Florida Bay also experienced a serious loss of seagrasses over a decade (1984–1994), which was partly
due to increased turbidity (Hall et al., 1999) and in
Chesapeake Bay losses or Zostera marina and Ruppia maritima were also related to increased turbidity as a result of eutrophication (Orth and Moore,
1983). In Waquoit Bay, Massachusetts Short and
Burdick (1996) related housing development and nitrogen loading to eelgrass habitat loss over the period
1987–1992 (Fig. 10). The effect occurred largely via
ground water and resulted in a gradual fragmentation
and loss of the meadows.
Examples of recolonization upon reduction of eutrophication are limited. The seagrass cover in Cockburn Sound, Western Australia, was markedly reduced between 1976 and 1981 as a response to
eutrophication, but reductions in nutrient loads in
the 1980s did not lead to recolonization (Walker
et al., Chapter 23). It is likely that alterations in
shelf-environments during the period without seagrasses have rendered the area unsuitable for seagrass growth (Kendrick et al., 2002). In contrast,
Posidonia coriaceae and Amphibolis griffithii have
recolonized former seagrass areas in Success Bank,
Western Australia, at surprizingly high rates in100
75
50
25
0
0
5000
10000
15000
20000 25000
Nitrogen loading (kg km -2 yr -1 )
Eelgrass area (%)
Log(y) = 1.648 - 0.000044x
R 2 = 0.888
1987
1988
1989
Fig. 10. Comparison of nitrogen loading rates and eelgrass area
in the Waquoit Bay estuary’s sub-basins over the first three years
of study. The log of eelgrass area is regressed against loading.
From Short and Burdick (1996), with permission from the Estuarine Research Federation.
volving both vegetative and sexual reproduction
(Kendrick et al., 1999; Walker et al., Chapter 23). An
extremely rapid eelgrass recolonization was also observed in the Archipelago of Southern Funen, Denmark. This area experienced an 80% reduction in
the distribution area of eelgrass following an anoxic
event during a warm summer period, but recovered
completely within 3 years through a combination of
vegetative growth of surviving shoots and germination of seeds (Rask et al., 2000). An even faster recolonization of Z. marina after anoxia-induced mortality was observed in the Thau Lagoon, French Mediterannean Sea (Plus et al., 2003).
Rapid recolonization seems possible if the disturbance causing the seagrass decline is limited in time
and space and if seedlings originating from the sediment bank or from neighbouring populations experience suitable growth conditions the following
year. By contrast, recovery of seagrass populations
from catastrophic decline on the landscape scale requires patch initiation from seeds transported from
distant populations and subsequent patch growth.
The survival chances of these initial patch stages
are low, and the formation of new extended patches
may, therefore, be a protracted process. Simulation
models show that small species with large recolonization potentials may recover within a few years
after a disturbance, while large species with small
recolonization potentials may require centuries to recover if the process is at all reversible (Duarte, 1995).
Colonization may be further delayed or impeded by
negative cascading effects (Duarte, 1995).
In many cases, declines of seagrass meadows are
not detected before marked losses have occurred