374
Kenneth A. Moore and Frederick T. Short
most development; overall, 60% of the Z. marina was
lost from this estuary in five years. In Ninigret Pond,
Z. marina distributions were compared over a 32year period using historical and recent maps; areal
distribution of Z. marina declined by 41% (Short
et al., 1996).
The seagrass decline in Chesapeake Bay has been
attributed to the combined effects of cumulative anthropogenic stresses over many decades and a large
tropical storm (Agnes) in 1972 that caused a pulsed
input of nutrients and sediment from the watershed
into the estuary (Kemp et al., 1983; Orth and Moore,
1983a, 1984). In addition to changing distributions
in areas where Z. marina has not died out completely,
the outer depth limits of the seagrass in the bay have
decreased approximately 0.5 m over time. In one major sub-estuary, the York River, historically (1950s)
the depth limits of Z. marina ranged from approximately 2.0 m (mean low water) at the mouth of the
estuary, to 1.5 m in the middle estuary and 1.0 m
at its former upriver limits of distribution (Moore
et al., 2000). Current depth limits at these locations
are approximately 1.5 m, 1.0 m and 0.0 m (mean
low water) respectively. This change in depth limits
suggest a decrease in light availability either through
increased water column turbidity or epiphyte accumulation. Complete loss in upriver areas suggests
that other factors may be limiting seagrass growth at
the shallowest depths. Moore et al. (1997) found that
pulsed turbidity levels in the spring correlated with
the dieback of transplants in this region. These increases in turbidity could be due to physical factors
including sediment resuspension (Koch, 1999b) or
increased watershed inputs. Other physical factors
associated with low tides over these flats (Koch and
Beer, 1996) may also contribute to the lack of regrowth of propagules in these areas.
Dredging, filling, marina development, boat activity, fishing practices, hardening of the shoreline
and anthropogenic nutrient and sediment discharge
all continue to impact Zostera and other seagrass
habitats and areas where they could return (Short
and Wyllie-Echeverria, 1996; Burdick and Short,
1999; Kendrick et al., 2000). In Maquoit Bay, Maine,
USA, dragging for mussels in 1999 created a 31.8 ha
bare area in the center of a large Z. marina meadow
(Neckles et al., 2005). Similar direct impacts to
Zostera beds have been documented elsewhere including clam dredging in Maryland and Virginia,
USA (Orth et al., 2002), mussel and cockle harvesting in the Dutch Wadden Sea (De Jonge and De
Jong, 1992), and scallop dredging (Fonseca et al.,
1984) and clam harvesting (Peterson et al., 1983)
in North Carolina, USA. Recovery times have not
been well documented, although in Maryland and
Virginia the recovery time for Z. marina beds exceeded three years (Orth et al., 2002) and in Maine,
complete recovery from dragging was predicted to
take over 17 years (Neckles et al., 2005). Bare areas
created by the dredging were also reported to experience increased erosion due to ray foraging activities
and other physical forces (Fonseca et al., 1984; Orth
et al., 2002) that may have inhibited recovery. Similarly, recovery of Z. capricorni from human impacts
in Moreton Bay, Australia was slowed by low levels
of grazing by dugongs (Preen, 1995).
Shellfish farming as well as other sources of excessive organic enrichment have resulted in both
anoxia and sediment reducing conditions that impact Zostera beds (De Casabianca et al., 1997;
Flindt et al., 1997; Terrados et al., 1999). Not only
do hypoxic/anoxic events affect faunal communities in these vegetated habitats (Guerrini et al.,
1999), but anaerobiosis as well as accompanying elevated sediment sulfide levels affect Zostera growth
and survival through effects on root metabolism
(Smith et al., 1988), nutrient uptake (Pregnall et al.,
1984) and photosynthetic processes (Goodman et al.,
1995).
B. Use of Zostera Species as an Indicator
The establishment of relationships between light
availability, water quality conditions and depth distribution of a widely distributed species such as
Z. marina (Dennison and Alberte, 1985; Duarte,
1991, Olesen, 1996) has provided an important
tool for establishing habitat requirements for the
species, and subsequently using the distribution or
presence/absence of the species as an indicator of
environmental conditions or health of a system
(Dennison et al., 1993; Short et al., 1993a). In the
Chesapeake Bay region, for example, habitat requirements for polyhaline regions of the Bay (Table
2) were established based upon seasonal medians
of water quality constituents (light attenuation, total
suspended solids, chlorophyll a, dissolved inorganic
nitrogen and dissolved inorganic phosphorus) that
characterized the local environments of areas with
either fluctuating or persistent beds of Z. marina or
the survival of Z. marina transplants (Batiuk et al.,
1992). Subsequently, water clarity requirements for
Kenneth A. Moore and Frederick T. Short
most development; overall, 60% of the Z. marina was
lost from this estuary in five years. In Ninigret Pond,
Z. marina distributions were compared over a 32year period using historical and recent maps; areal
distribution of Z. marina declined by 41% (Short
et al., 1996).
The seagrass decline in Chesapeake Bay has been
attributed to the combined effects of cumulative anthropogenic stresses over many decades and a large
tropical storm (Agnes) in 1972 that caused a pulsed
input of nutrients and sediment from the watershed
into the estuary (Kemp et al., 1983; Orth and Moore,
1983a, 1984). In addition to changing distributions
in areas where Z. marina has not died out completely,
the outer depth limits of the seagrass in the bay have
decreased approximately 0.5 m over time. In one major sub-estuary, the York River, historically (1950s)
the depth limits of Z. marina ranged from approximately 2.0 m (mean low water) at the mouth of the
estuary, to 1.5 m in the middle estuary and 1.0 m
at its former upriver limits of distribution (Moore
et al., 2000). Current depth limits at these locations
are approximately 1.5 m, 1.0 m and 0.0 m (mean
low water) respectively. This change in depth limits
suggest a decrease in light availability either through
increased water column turbidity or epiphyte accumulation. Complete loss in upriver areas suggests
that other factors may be limiting seagrass growth at
the shallowest depths. Moore et al. (1997) found that
pulsed turbidity levels in the spring correlated with
the dieback of transplants in this region. These increases in turbidity could be due to physical factors
including sediment resuspension (Koch, 1999b) or
increased watershed inputs. Other physical factors
associated with low tides over these flats (Koch and
Beer, 1996) may also contribute to the lack of regrowth of propagules in these areas.
Dredging, filling, marina development, boat activity, fishing practices, hardening of the shoreline
and anthropogenic nutrient and sediment discharge
all continue to impact Zostera and other seagrass
habitats and areas where they could return (Short
and Wyllie-Echeverria, 1996; Burdick and Short,
1999; Kendrick et al., 2000). In Maquoit Bay, Maine,
USA, dragging for mussels in 1999 created a 31.8 ha
bare area in the center of a large Z. marina meadow
(Neckles et al., 2005). Similar direct impacts to
Zostera beds have been documented elsewhere including clam dredging in Maryland and Virginia,
USA (Orth et al., 2002), mussel and cockle harvesting in the Dutch Wadden Sea (De Jonge and De
Jong, 1992), and scallop dredging (Fonseca et al.,
1984) and clam harvesting (Peterson et al., 1983)
in North Carolina, USA. Recovery times have not
been well documented, although in Maryland and
Virginia the recovery time for Z. marina beds exceeded three years (Orth et al., 2002) and in Maine,
complete recovery from dragging was predicted to
take over 17 years (Neckles et al., 2005). Bare areas
created by the dredging were also reported to experience increased erosion due to ray foraging activities
and other physical forces (Fonseca et al., 1984; Orth
et al., 2002) that may have inhibited recovery. Similarly, recovery of Z. capricorni from human impacts
in Moreton Bay, Australia was slowed by low levels
of grazing by dugongs (Preen, 1995).
Shellfish farming as well as other sources of excessive organic enrichment have resulted in both
anoxia and sediment reducing conditions that impact Zostera beds (De Casabianca et al., 1997;
Flindt et al., 1997; Terrados et al., 1999). Not only
do hypoxic/anoxic events affect faunal communities in these vegetated habitats (Guerrini et al.,
1999), but anaerobiosis as well as accompanying elevated sediment sulfide levels affect Zostera growth
and survival through effects on root metabolism
(Smith et al., 1988), nutrient uptake (Pregnall et al.,
1984) and photosynthetic processes (Goodman et al.,
1995).
B. Use of Zostera Species as an Indicator
The establishment of relationships between light
availability, water quality conditions and depth distribution of a widely distributed species such as
Z. marina (Dennison and Alberte, 1985; Duarte,
1991, Olesen, 1996) has provided an important
tool for establishing habitat requirements for the
species, and subsequently using the distribution or
presence/absence of the species as an indicator of
environmental conditions or health of a system
(Dennison et al., 1993; Short et al., 1993a). In the
Chesapeake Bay region, for example, habitat requirements for polyhaline regions of the Bay (Table
2) were established based upon seasonal medians
of water quality constituents (light attenuation, total
suspended solids, chlorophyll a, dissolved inorganic
nitrogen and dissolved inorganic phosphorus) that
characterized the local environments of areas with
either fluctuating or persistent beds of Z. marina or
the survival of Z. marina transplants (Batiuk et al.,
1992). Subsequently, water clarity requirements for