2012 NORTHEASTERN NATURALIST 19(3):501–516
Abundance and Species Composition Surveys of
Macroalgal Blooms in Rhode Island Salt Marshes
Christine Newton1, 2,* and Carol Thornber2
Abstract - Excessive growth of macroalgae in estuarine systems is becoming increasingly
common among coastal communities throughout the world. Despite repeated
observations of macroalgae growing or deposited among the stems of lower marsh plants,
few studies have quantitatively documented the presence of macroalgae in salt marsh
communities. We conducted monthly surveys during 2009 and 2010 to document the
species composition and abundance of the macroalgal community, along with associated
biological and physical parameters, in 9 Rhode Island salt marshes. Macroalgae were
found in every site during each month sampled, with a peak biomass during the fall,
reaching densities up to 1500 g/m2 (wet mass). Nearly 80% of the macroalgae was found
in the fi rst 2 m of the lower marsh zone. Fucus spp. were dominant throughout the year,
accounting for almost 70% of the annual abundance. While several biological parameters
were measured in this study that may contribute to macroalgal accumulation, it is likely
that a combination of biotic and abiotic factors drive macroalgal accumulation patterns
in these systems.
Eutrophication poses an important threat to coastal systems (Andersen
et al. 2006, Cloern 2001, Lee and Olsen 1985, Nixon et al. 1995, Scavia and
Bricker 2006, Valiela et al. 1997). Increases in nutrient levels can facilitate the
extremely rapid growth of macroalgae and can result in the formation of macroalgal
blooms (Granger et al. 2000, McGlathery 2001, Nixon and Buckley
2002, RI DEM 2003, Rosenberg 1985). Blooms can be particularly extensive
in areas with low water exchange, such as estuaries, and in systems that experience
changes in temperature and light levels (Lee and Olsen 1985, Taylor et
al. 1995). While macroalgae provide food, habitat, and refuge for many marine
animals, severe ecological and economic problems have been associated with
macroalgal blooms in coastal regions (Granger et al. 2000, Hauxwell et al.
1998, Liu et al. 2009, Rosenberg 1985, Sfriso et al. 1992, Thomsen and Mc-
Glathery 2006, Valiela et al. 1997).
In central and southern New England, macroalgal blooms have been studied
in detail in estuarine and open-coast environments, where they are comprised
primarily of Ulva spp., Cladophora spp., Gracilaria spp., and Polysiphonia/
Neosiphonia spp. (Lyons et al. 2009; Valiela et al. 1992, 1997; C.S. Thornber
and M. Guidone, University of Rhode Island, Kingston, RI, unpubl. data). Mats
1Northeastern University, Marine Science Center, 430 Nahant Road, Nahant, MA 01908.
2Department of Biological Sciences, College of Environmental Life Sciences, University
of Rhode Island, 120 Flagg Road, Kingston, RI 02881. *Corresponding author -
502 Northeastern Naturalist Vol. 19, No. 3
of bloom macroalgae can also drift into nearby salt marshes, becoming entangled
among cordgrass stalks and accumulating on intertidal flats (Boyer and Fong
2005; C. Newton, pers. observ.). However, there are relatively few fi eld observations
of the presence of macroalgae in salt marsh habitats. These reports have
been species specifi c and geographically scattered (Brinkhuis 1977a, b; Dawes
1998; Thomsen et al. 2009; Udell et al. 1969). The biomass and distribution of
perennial macroalgal species has received careful study in northern New England
marshes (e.g., Chock and Mathieson 1983, Mathieson and Penniman 1986,
Mathieson et al. 2006) where ecads of perennial macroalgal species can be seen
growing among the stems of lower marsh plants (Chapman and Chapman 1999,
Gerard 1999). To the knowledge of the authors, similar studies have not been
conducted in southern New England regions.
Without a foundational knowledge of the relative abundance and diversity
of macroalgal species in southern New England salt marsh habitats, we cannot
explore potential ecological impacts that bloom- and non-bloom-forming
macroalgal presence may have on an ecosystem. For example, in the laboratory,
Boyer and Fong (2005) found evidence for a bloom-macroalgal-mediated link
in the transfer of nutrients from the water column to the marsh plant Salicornia
virginica L. (Glasswort). In addition, facilitative interactions have been documented
in the fi eld between the ecad scorpioides of the perennial brown alga
Ascophyllum nodosum ecad scorpioides Hauck and Spartina alterniflora Loisel
(Smooth Cordgrass), the dominant native lower marsh plant along western Atlantic
coastlines (Gerard 1999). Given the potential ecological roles macroalgae
could play, we sought to examine: (1) the extent of macroalgae present in these
salt marsh ecosystems, (2) the dominant species of macroalgae found in these
salt marshes, (3) how high in the marsh could macroalgae potentially impact
community interactions, and (4) which biological factors were associated with
macroalgal accumulation. We investigated these questions in a suite of southern
New England salt marshes located on and around Narragansett Bay, RI. We discuss
our results in the context of potential ecological impacts macroalgal blooms
may have on salt marsh community structure and function.
We conducted monthly surveys at 6 salt marshes within Narragansett Bay,
RI, as well as 3 located in Rhode Island Sound (fig. 1). Sites were selected
to encompass an inner- to outer-bay gradient, as well as a eutrophic gradient
(Granger et al. 2000, Oviatt 2008) between an outer coastal site at Point Judith
Pond and Greenwich Bay, a highly eutrophic region of northern Narragansett
Bay with frequent macroalgal blooms (Guidone et al. 2010). We selected our
study sites to cover a wide range of salt marsh habitats, ranging from fringing
marshes (e.g., Rocky Hill and Sheffield Cove) to large protected sanctuaries
(e.g., Marsh Meadows and Galilee Bird Sanctuary). Each of these marshes exhibited
typical banding zonation patterns, with the seaward borders dominated
2012 C. Newton and C. Thornber 503
by the native cordgrass S. alterniflora (short and tall form; Golet et al., in
press; Nixon and Oviatt 1973). The mid-marsh was characterized by a variety
of plants including Spartina patens (Aiton) Muhl. (Salt Meadow Grass), but
the high marsh terrestrial borders varied at each site. However, these midand
high-marsh zones are rarely inundated, rarely contained macroalgae, and
therefore were not the focus of this study.
Surveys were conducted monthly during spring low tides from May 2009
through August 2010. At each site, we placed two 10-m transects (hereafter
figure 1. Salt marsh survey sites encompassing a gradient from the inner bay to the outer
coast of Narragansett Bay, R.I. Circle size indicates the mean summer abundance of
macroalgae relative to other study sites.
504 Northeastern Naturalist Vol. 19, No. 3
referred to as “edge” transects) along the lower edge of the regularly (daily)
flooded S. alterniflora zone, where it borders open water or mudflats, to determine
the species composition and biomass of macroalgae that regularly
impact low-zone S. alterniflora. The two transects were placed more than 5 m
apart, except where restricted by marsh topography. Ten 0.0625-m2 quadrats,
subdivided into 25 squares (to ensure accuracy), were placed at 1-m intervals
along each transect. A third transect (hereafter referred to as the “vertical”
transect) was placed perpendicular to these transects to determine how high in
the marsh drift macroalgae typically occur. A 0.0625-m2 quadrat was placed at
1-m intervals along this 10-m transect until it was evident that there were no
algae present farther up in the marsh.
Within each quadrat, we recorded the percent cover of all macroalgae
before collecting it in a Ziploc bag for further analysis in the laboratory.
Subsequently, we spun the macroalgae from each quadrat 15 times in a salad
spinner to remove excess moisture and then measured the wet mass (hereafter,
WM) of each genus/species. The removal of algae during monthly samples
did not impact the following month’s survey because a) we varied the location
of each transect slightly among months, and b) drift macroalgae can enter the
marsh during each tidal cycle.
In the fi eld, we also measured biotic factors that might impact macroalgal
accumulation, such as the percent cover of living S. alterniflora, exposed roots,
and dead S. alterniflora, as well as the numbers of Littorina spp. (periwinkles),
mussels (Geukensia demissa Dillwyn [Ribbed Mussel] and Mytilus edulis L.
[Blue Mussel]), and crabs (primarily Hemigrapsus sanguineous De Haan [Asian
Shore Crab] and Carcinus maenas L. [Green Crab]) present within each quadrat.
Additionally, as abiotic factors such as ice cover and water flow might impact
macroalgal accumulation or growth (Martins et al. 2001), we measured the percent
ice cover during winter. We determined relative flow rates using modifi ed
clods, constructed with wooden stakes and plaster cubes, after Doty (1971) and
River and Edmunds (2001). Eleven clods per site were deployed at each site for
72 hours. The dissolution of plaster over this period provided a measurement of
the relative water flow rates for each site.
We analyzed spatial and temporal variation in the overall abundance of drift
macroalgae using two-way ANOVAs. Tukey post hoc tests were run on any
signifi cant factors. A graphical representation of species composition patterns
was created using non-metric multi-dimensional scaling (nMDS) ordinations,
following the calculation of Bray-Curtis similarity matrices on square-root
transformed data. Biological factors were overlaid on the nMDS plot as vectors
to indicate potential relationships between biological variables and species
composition patterns. We used an analysis of similarity (ANOSIM) to test
multivariate data (species composition) for further differences among sites and
seasons. A combination of linear regressions and one-way ANOVAs were used to
determine if macroalgal accumulation was correlated with the biotic factors and
clod-dissolution rates measured in this study. Univariate statistics were run using
2012 C. Newton and C. Thornber 505
JMP 8.0 (SAS Institute, Cary, NC), while multivariate analyses were run using
PRIMER-E v.6 (www.primer-e.com).
We found macroalgae in every salt marsh site during every month surveyed.
In our “edge” transects, the abundance of macroalgae ranged from 0.1 g/m2
WM (September 2009, Bissel Cove) to 1487.0 g/m2 WM (July 2009, Sheffield
Cove), with an overall mean of 84.8 g/m2 WM. Macroalgal abundance
was typically 1.5 times higher during the summer and fall than in the winter or
spring (Two-factor ANOVA: F3, 2429 = 7.78, P less than 0.001; Table 1). Among sites,
macroalgal abundance at Sheffield Cove was nearly twice as high as Bluff
Hill Cove (325g/m2 WM vs. 180g/m2 WM) and ten times higher than the average
of the remaining seven sites (Two-factor ANOVA: F8, 2429 = 154.96, P less than
0.001; fig. 1), with a significant season*site interaction (Two-factor ANOVA:
F24, 2429 = 17.18; P < 0.001; Table 1).
We found a total of 26 genera of macroalgae along the lower edge of the
nine salt marshes examined (Table 2). Fucus spp., primarily F. vesiculosus and
F. spiralis, accounted for 68.5% of the annual macroalgal abundance in these
New England marshes. The second most dominant species/genera varied seasonally
(fig. 2). During the spring, macroalgal biomass was dominated by the
filamentous brown alga Pylaiella littoralis and tubular Ulva spp. There was a
wide range of species present during the summer months, including both
blade and tubular Ulva spp., P. littoralis, and Rhizoclonium spp. During the
fall season, Codium fragile subsp. fragile and A. nodosum dominated, while
Gracilaria spp., A. nodosum, and Polysiphonia/Neosiphonia spp. prevailed
during the winter.
Species composition varied signifi cantly among sites (ANOSIM: R = 0.773,
P = 0.034). Bluff Hill Cove and Sheffi eld Cove, despite being geographically
Table 1. Site and seasonal differences in mean macroalgal abundance. Sites are arranged from inner-
to outer-bay gradient. Data are presented as mean wet mass (g/m2 ± 1 S.E.).
Spring Summer Fall Winter
Overall 57.45 (4.01) 89.81 (5.47) 113.08 (10.11) 78.56 (6.94)
Chepiwanoxet 6.45 (2.01) 53.89 (10.07) 13.42 (5.32) 175.42 (35.33)
Rocky Hill 13.31 (2.16) 9.34 (1.62) 7.06 (2.12) 26.61 (6.69)
Bissel Cove 56.32 (9.03) 67.88 (13.21) 21.74 (15.55) 4.12 (2.23)
Marsh Meadows 54.04 (7.72) 27.41 (6.66) 31.51 (11.81) 17.52 (2.75)
Sheffi eld Cove 176.87 (27.65) 375.39 (22.76) 411.88 (38.27) 277.69 (22.61)
Fox Hill 72.51 (16.31) 15.34 (3.14) 10.18 (5.12) 5.72 (1.46)
Narrow River 63.74 (11.74) 33.00 (6.65) 10.69 (3.17) 21.84 (4.94)
Bluff Hill Cove 60.18 (7.24) 209.87 (22.12) 335.31 (23.80) 147.14 (17.42)
Galilee 83.24 (13.37) 28.87 (7.19) 8.68 (6.38) 4.77 (1.47)
506 Northeastern Naturalist Vol. 19, No. 3
distinct, grouped together on the lower left of the nMDS plot (fig. 3), indicating
they had similar macroalgal abundances and species compositions. Minimal
evidence of an inner- to outer-bay eutrophic gradient was found, although the 2
sites located in upper Narragansett Bay (Greenwich Bay), grouped together on
the right of the nMDS plot with similar abundances and species compositions
(Chepiwanoxet and Rocky Hill; fig. 3). In contrast, species composition did not
vary signifi cantly among seasons (ANOSIM: R = 0.526, P = 0.122).
Vertical distance of macroalgal accumulation
We found that macroalgae extended 10 m or more from the lower S. alternifl
ora edge and on occasion entered the mid-marsh (i.e., S. patens) zone. However,
Table 2. Macroalgal species/genera present during monthly surveys in Rhode Island salt marshes
from May 2009 through August 2010. Aindicates genus is currently under taxonomic revision in
Rhode Island; Bas identifi ed in Schneider (2010)
Chlorophyta (Green Algae)
Chaetomorpha spp. (green thread)
Cladophora spp. (green tuft)
Codium fragile (Suringar) Hariot subsp. fragile (Suringar) Hariot (Dead Man’s fingers)
Gayralia oxysperma (Kützing) K.L. Vinogradova ex Scagel et al.
Monostroma grevillei (Thuret) Wittrock (Green Laver)
AUlva blade spp. (sea lettuce)
AUlva tube spp. (green nori)
Heterokontophyta: Phaeophyceae (Brown Algae)
Ascophyllum nodosum ecad scorpioides Hauck (Knotted Wrack)
Desmarestia spp. (stink weed or sea sorrel)
Fucus distichus (Rockweed)
Fucus spiralis (Spiral Wrack or Flat Wrack)
Fucus vesiculosus (Bladderwrack or Rockweed)
Petalonia fascia (O.F. Müller) Kuntze (Mini Kelp, Sea Petals, or Leaf Weed)
Pylaiella littoralis (L.) Kjellman
Sargassum fi lipendula C. Agardh (Sargasso Weed)
Scytosiphon lomentaria (Lyngbye) Link (Sausage Weed)
Rhodophyta (Red Algae)
Agardhiella subulata (C. Agardh) Kraft & M.J Wynne (Agardh’s Red Weed)
Antithamnion spp. (red sea skein)
Ceramium spp. (banded weed)
Champia parvula (C. Agardh) Harvey (Barrel Weed)
Chondrus crispus Stackhouse (Irish Moss)
Cystoclonium purpureum (Hudson) Batters (Grapevine Weed)
Gracilaria tikvahiae McLachlan (Red Spaghetti)
Gracilaria vermiculophylla (Ohmi) Papenfuss (Red Spaghetti or Graceful Red Weed)
Grateloupia turuturu Yamada (Devil’s Tongue Weed)
B“Heterosiphonia” japonica Yendo
Neosiphonia harveyi (J.W. Bailey) M.-S. Kim, H.-G. Choi, Guiry & G.W. Saunders
(Harvey’s Siphon Weed)
Polysiphonia spp. (polly or tubed weeds)
Porphyra spp. (nori or laver)
2012 C. Newton and C. Thornber 507
on average, 80% of the total biomass from the vertical transect was found within
the fi rst 2 m of the S. alterniflora zone (fig. 4).
figure 3. nMDS plot of macroalgal community compositions from May 2009 and August
2010. Biological factors measured at each site are overlaid as vectors.
figure 2. Seasonality of the eight most abundant macroalgal species along the lower
S. alterniflora edge (representing 96% of total abundance). Bars are mean wet mass ± 1
508 Northeastern Naturalist Vol. 19, No. 3
Biotic and abiotic factors driving macroalgal accumulation
Site characteristic vectors (of biological factors) that overlay the multivariate
nMDS plot (fig. 3) indicate that the species composition in eutrophic areas
(Chepiwanoxet and Rocky Hill) is correlated with mussel density (G. demissa
and M. edulis). By contrast, the species composition at Bluff Hill Cove and Sheffi
eld Cove, although geographically distinct, appeared to be correlated by the
overall percent cover of macroalgae along the lower S. alterniflora edge, where,
on average, 55% and 83%, respectively, of the plots were covered with macroalgae.
This corresponds to our abundance surveys, where signifi cantly higher
abundances were found at these two sites (P < 0.001; fig. 1).
Linear regressions for overall macroalgal abundance and each of the
biological factors measured (% cover of living, dead, and exposed roots of
S. alterniflora; % of bare areas; ice cover; and the number of mussels, crabs,
or Littorina spp.) did not reveal significant relationships (R2 < 0.05 for all
factors). Additionally, while the sites in eutrophic areas (Chepiwanoxet and
Rocky Hill) had 1.5 times higher flow rates than the remaining sites (F8, 88 =
7.63, P < 0.001), the patterns of total macroalgal accumulation could not be
explained by this factor alone (R2 < 0.01). However, water flow rates were correlated
with mussel densities (R2 = 0.62), which had been shown previously to
correlate with macroalgal species composition.
figure 4. Vertical distance of macroalgal biomass accumulation from the lower Spartina
alterniflora edge. Data are mean wet mass ± 1 standard error.
2012 C. Newton and C. Thornber 509
Our study provides the fi rst comprehensive record, to our knowledge, of the
abundance and species composition of macroalgal bloom- and non-bloom-forming
species in Rhode Island salt marshes. The abundance of macroalgae found in
our surveys showed signifi cant temporal variability, with the highest accumulations
during the summer and fall. Increases in abundance were typically driven
by an overall increased biomass of the perennial F. distichus, F. spiralis, and
F. vesiculosus, coupled with increased bloom-forming Ulva spp. and Pylaiella
littoralis during the summer and the perennial C. fragile subsp. fragile during
fall. Seasonality is to be expected in the temperate New England climate, where
summer brings warmer temperatures, along with increased light and nutrient
availability, which facilitate macroalgal growth in Narragansett Bay (Karentz
and Smayda 1984, RI DEM 2003). Interestingly, the maximum abundance was
reached during fall, which may be due to the seasonal cycles seen in New England
salt marsh communities. Unlike marshes in southern latitudes, northern
marshes are limited in lower marsh development, with plants only growing from
May through September. In the fall, aboveground plant material is broken down
into detritus that forms large wrack mats, providing a natural pulse of nutrients
and organic matter from the decomposing wrack during the fall months (Bertness
1999, Pennings and Bertness 2001, Valiela and Rietsma 1995). Additional
nutrients from this decomposition may also facilitate further macroalgal tissue
growth in the fall (Lee and Olsen 1985, Taylor et al. 1995, Teal 1962).
While Chock and Mathieson (1983) found similar seasonal trends of fall
peaks in perennial ecad species in a New Hampshire salt marsh, the open
ocean beaches and rocky shores in our study area experience peak macroalgal
abundance during June and July, with up to 920 g/m2 (C. Thornber, unpubl.
data). While the majority of macroalgae found in salt marshes were perennial
fucoid species (Fucus spp. and A. nodosum) found attached to substrata,
a large portion of the macroalgae we found formed drifting material that accumulated
on intertidal flats or became entangled in S. alterniflora stalks
(C. Newton, pers. observ.). It is likely that these macroalgae were transported
from other sources, such as nearby beaches and rocky shores where the peak
abundance of macroalgal blooms occurs during summer (C. Thornber, unpubl.
data). As drift macroalgae enters the water column from these nearby systems,
they may drift into nearby salt marsh habitats due to currents and tidal cycles
over the course of the summer, becoming trapped by topography, low water
flow, or reduced flushing rates experienced in salt marsh systems.
A total of 26 different genera were found during our surveys, with representatives
from each of the three groups of macroalgae: Chlorophyta,
Heterokontophyta: Phaeophyceae, and Rhodophyta (Table 2). The high algal
diversity we documented contrasts with the results of previous distributional
surveys that have been conducted. For example, Brinkhuis (1976), Chock and
Mathieson (1983), and Roman et al. (1990) focused their studies on fucoid
510 Northeastern Naturalist Vol. 19, No. 3
abundance and productivity. While Roman et al. (1990) did note the other species
found in Nauset Marsh (MA), only six species were found intertidally, compared
to our 26 genera of macroalgae. Biodiversity may be highly site-specifi c; we
found species richness ranged from a minimum of 8 species up to 18 species per
site (mean = 13.3 species per site). Therefore, the increased biodiversity may be
due to the nature of our study in which we examined 8 marshes compared to the
single marsh in Roman et al. (1990). As biodiversity can have important consequences
for ecosystem functioning (e.g., Duffy 2002, Duffy et al. 2007, Schwartz
et al. 2000, Worm and Duffy 2003, Worm et al. 2006), understanding the range of
macroalgal biodiversity in salt marsh habitats may yield insights into the ecological
dynamics of these systems.
Despite the wide variety of species encountered, we did not find significant
seasonal differences in species composition, most likely due to the high abundance
of fucoids (Fucus spp. and A. nodosum) across all seasons (74.9% of the
total abundance). Fucoids were less dominant in our surveys than in research
conducted by Chock and Mathieson (1983), who found that they accounted
for 98% of the total abundance of macroalgal abundance in Cedar Point, Little
Bay, NH. However, our intertidal surveys showed a higher total species richness
(n = 26) than in Cedar Point (n = 20), which may account for the lower
abundance of fucoids.
We found that signifi cant differences existed in species composition at the different
sites. The biological factors we measured in this study may explain some
of these differences and warrant further investigation. For example, it appears the
species composition at our northern Bay sites (Chepiwanoxet and Rocky Hill)
may be driven in part by high mussel densities, which may serve as attachment
points for macroalgal thalli (Hardwick-Witman 1985). However, the factors examined
in this study could not explain differences in overall macroalgal biomass
among sites. There are a number of other biotic and abiotic factors that may be
responsible for patterns of macroalgal abundance, such as nutrient levels, marsh
topography, competition, and herbivory. It is likely that a combination of these
factors, along with the biological and water-flow parameters measured in this
study, drive both macroalgal accumulation and species composition.
At biomass levels of up to 1500 g/m2 WM, macroalgal accumulations have the
potential to signifi cantly impact salt marsh community structure. Previous studies
have shown that these communities are extremely susceptible to small changes in
both bottom-up (i.e., an introduction of nutrients in to the system) and top-down
(i.e., introduction of predators) control over local food webs and trophic structures
(Denno et al. 2002, Hauxwell et al. 1998, Levine et al. 1998, Silliman and
Bertness 2002). Increased macroalgal abundance may act as a pulsed bottom-up
effect on these systems by increasing nutrient availability through their release
of nutrients during tissue decomposition (Boyer and Fong 2005). The increase in
algal abundance may also modify habitat structure and alter species interactions
by shading or smothering, resulting in a reduced canopy height (Alexander and
Robinson 2006, Sanger et al. 2004). Despite these potential impacts, relatively
2012 C. Newton and C. Thornber 511
little research has examined the effect of macroalgal presence on salt marsh community
structure (but see Boyer and Fong 2005; Chapman and Chapman 1999;
Chock and Mathieson 1983; Gerard 1999; Mathieson et al. 2006; C. Newton,
Recent efforts to implement nutrient reductions in Narragansett Bay may also
have a signifi cant effect on the macroalgal biomass and species composition
found throughout the Bay (Deacutis 2008, Nixon et al. 2008). Using stable isotopes,
previous studies have shown that sewage-derived nitrogen is present within
bloom-forming macroalgae found in Narragansett Bay (Oczkowski et al. 2008,
Thornber et al. 2008). Increased nitrogen removal from wastewater plants will
reduce the amount of macroalgae present within the Bay; however, it is likely to
take several years before signifi cant impacts are seen. For example, in Mumford
Cove, CT, mats of Ulva disappeared within two years of wastewater discharge
reductions (Vaudrey et al. 2002). Similarly, reductions of Ulva were seen in Mondego
estuary (Portugal) following nutrient diversions and increased circulation
throughout the estuary (Leston et al. 2008). Following these reductions of nutrients
and Ulva mats, the restoration of submerged aquatic vegetation is possible,
as seen in Mondego with the increase in Zostera noltii Hornem (Dwarf Eelgrass)
and in Mumford Cove with the return of eelgrass meadows and Zostera marina
L. (Common Eelgrass) within 10–15 years. A similar example is seen in Tampa
Bay, FL, where the biomass of rapidly growing seaweeds was reduced within 1–3
years following nutrient reductions (Johansson 2002). Restoration of submerged
aquatic vascular rooted grass beds was seen over the next 8–10 years in both
Tampa Bay and Sarasota Bay (FL). However, one of the greatest weaknesses to
tracking the effects of nutrient reductions is the lack of baseline data prior to the
nutrient changes. With our study, we are able to provide adequate data to properly
assess the impacts on macroalgae before and following nutrient reduction efforts
in Narragansett Bay.
In addition to eutrophication, salt marshes are particularly susceptible to
climate-related ecological changes. Rising temperatures, increasing sea level,
changes in nutrient availability and UV radiation, along with stronger and more
frequent storms, are predicted to signifi cantly impact estuarine communities
(Harley et al. 2006, IPCC 2007). Ecological changes such as these have the potential
to alter population dynamics, shift species distributions, increase the rate
of species extinctions, and change species interactions (fields et al. 1993, Lotze
and Worm 2002, Sanford 1999). Despite efforts to restore salt marsh habitats,
these areas are continuously facing new threats from climate change and require
our constant attention (Harley et al. 2006). The information from this study provides
an important baseline for understanding macroalgal bloom dynamics in an
economically and commercially valuable habitat. While each salt marsh is unique
in its physical characteristics and biological interactions, all marshes play an
important role as the transition area between land, brackish, and saline habitats
by stabilizing shorelines, acting as a buffer between the terrestrial and marine
interface, and providing nurseries and habitats to commercially valuable species
512 Northeastern Naturalist Vol. 19, No. 3
(Bertness 1999, Taylor et al. 1995, Teal 1962, Valiela 1983, Valiela et al. 2004).
With an understanding of the biodiversity and abundance of macroalgae in salt
marsh systems, it is now imperative to explore how macroalgae may be affecting
these valuable habitats, as any alterations to these fragile ecosystems have the
potential to signifi cantly change the interactions within salt marsh communities,
as well as interactions with adjacent ecosystems.
The authors would like to thank S. Rinehart, C. Blewett, A. Heinze, K. Hyman, N.
Millette, M. Nepshinsky, and E. Vincent for their assistance in the field and laboratory.
F. Golet, G. Kraemer, M. Guidone, E. Preisser, N. Rohr, and two anonymous
reviewers have provided valuable comments towards this research. Also, thank you to
J. Ramsay for assisting with figures. Access to field sites was provided by the Rhode
Island Department of Environmental Management, the Rhode Island Audubon Society,
and the Rocky Hill School (E. Greenwich, RI). Funding for this research was provided
by NOAA grant #NA09NMF4720259: The Narragansett Bay Window Program 2009,
Quebec-Labrador Foundation Sounds Conservancy, the University of Rhode Island,
and Rhode Island Sea Grant.
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