August 6th 2013–Assemblyman Goldfeder inspects the just completed Rulers Bar Wetland Island Project in Jamaica Bay. The assemblyman has been a big supporter of Jamaica Bay since his first days in office and it was great to see him take the time to travel out to the island in the middle of jamaica bay and see the results first hand!!
Dan Mundy of Jamaica Bay Ecowatchers and Assemblyman Goldfeder discuss the next phase of restoration projects in Jamaica Bayby
June 2 2013
Great progress today in the effort to restore Rulers Bar Wetlands island with over 160 volunteers !! This is a coordinated effort between the Jamaica Bay Ecowatchers and the American Littoral Society. Thanks to the team of volunteers who came out and to the boat captains of the six vessels that ferried the large group out to the island. This was no small feat and took alot of coordination and hard work. No doubt it appears worth it as the hundred and sixty volunteers planted thousands of plugs before the rising tides forced us off the island !! A great day was had by all!!by
Coastal eutrophication as a driver of salt marsh loss
Linda A. Deegan1, David Samuel Johnson1,2, R. Scott Warren3, Bruce J. Peterson1, John W. Fleeger4, Sergio Fagherazzi5
& Wilfred M.Wollheim6
Salt marshes are highly productive coastal wetlands that provide
important ecosystem services such as storm protection for coastal
cities, nutrient removal and carbon sequestration. Despite protective
measures, however, worldwide losses of these ecosystems have
accelerated in recent decades1. Here we present data from a nineyear
whole-ecosystem nutrient-enrichment experiment. Our study
demonstrates that nutrient enrichment, a global problem for
coastal ecosystems2–4, can be a driver of salt marsh loss. We show
that nutrient levels commonly associated with coastal eutrophication
increased above-ground leaf biomass, decreased the dense,
below-ground biomass of bank-stabilizing roots, and increased
microbial decomposition of organic matter. Alterations in these
key ecosystem properties reduced geomorphic stability, resulting
in creek-bank collapse with significant areas of creek-bank marsh
converted to unvegetated mud. This pattern of marsh loss parallels
observations for anthropogenically nutrient-enriched marshes
worldwide, with creek-edge and bay-edge marsh evolving into mudflats
and wider creeks5–7. Our work suggests that current nutrient
loading rates to many coastal ecosystems have overwhelmed the
capacity of marshes to remove nitrogen without deleterious effects.
Projected increases in nitrogen flux to the coast, related to increased
fertilizer use required to feed an expanding human population,may
rapidly result in a coastal landscape with less marsh, which would
reduce the capacity of coastal regions to provide important ecological
and economic services.
An accelerated global nitrogen cycle1–3 has greatly increased the flow
of reactive nitrogen (primarily as NO3
2) from land to coastal marine
ecosystems, causing harmful algal blooms, hypoxia and fisheries
losses5,8. Salt marshes occupy a critical interface between the land
and the sea, where they provide important ecological and economic
services, such as nutrient removal, storm protection for coastal cities
and carbon sequestration, and habitats for numerous species of fish,
birds and invertebrates. It is thought that salt marshes can protect
coastal bays by removing land-derived nutrients9,10, a conclusion based
on measures of whole-system nutrient budgets11,12 and plot-level
experiments in which added nutrients were transformed into greater
above-ground plant production (primarily cordgrass Spartina spp.) or
denitrified4. Globally between a quarter and half of the area of the
world’s tidal marshes has been lost, and although multiple factors (sealevel
rise, development, loss of sediment supply) are known to contribute
to marsh loss1, in some locations the drivers remain unexplained.
Understanding the mechanisms underlying the continued loss of this
ecologically and economically important ecosystem is a global priority.
Here we present an ecosystem-level experimental approach to
understanding how the intertwined responses of plant biomass allocation,
microbial decomposition, and geomorphic stability to coastal
nutrient enrichment may drive salt marsh loss. For nine years (2004–
2012) we have enriched multiple whole-ecosystemmarsh landscapes to
nutrient levels that correspond to moderately-to-highly eutrophic waters
by adding dissolved nutrients to flooding tidal water13. Approximately
50%–60% of the added NO3
2 was processed (assimilated or denitrified)
in the nutrient-enriched systems; the remainder was exported in
ebbing tidal water12. The large scale of this experiment, which included
creeks, mudflats, tall-form smooth cordgrass (Spartina alterniflora) at
the creek-channel edge and saltmeadow cordgrass (S. patens) in the
high marsh, has revealed interactions that would not be apparent from
plot-level experiments in individual habitats.
Nutrient enrichment may invoke a series of positive feedbacks by
altering ecosystem processes that affect below-ground dynamics and
creek-bank stability, leaving marshes more susceptible to the erosive
forces of storms and sea-level rise and gravitational slumping. In less
than a decade, a cascade of changes induced by nutrient enrichment
resulted in loss of low marsh along the creek-bank edge (Fig. 1a–f) and
a corresponding loss of ecosystem function. Smooth cordgrass along
the creek-bank edge responded to nutrient enrichment with increased
above-ground biomass expressed as heavier, taller shoots (Fig. 2a),
lower structural compounds (decrease of about half in foliar lignin),
and increased N content (Table 1), with response ratios comparable to
plot-level nutrient-enrichment experiments4,14. Increased plant height
coupledwith less structural tissue causedmore extensive areas of smooth
cordgrass to fall over (lodge)—awell-known response to over-fertilizing
grasses15.Using permanent transects and high-precision global positioning
system (GPS) mapping across the elevation gradient, we found no
evidence (D.S.J., R.S.W. and L.A.D., manuscript in preparation) for the
hypothesized shift in the up-elevation boundary between S. alterniflora
and S. patens in response to nutrients15. In nutrient-enriched marshes,
smooth cordgrass allocated less photosynthate to nutrient-gathering
roots and storage rhizomes, resulting in a third less total below-ground
biomass and a lower root:shoot ratio (Fig. 2b, c). Two smooth cordgrass
growth attributes, a highly plastic above-ground/below-ground allocation16
and foliar uptake ofNO3
2(ref. 17), contribute to the reductions in
total below-ground biomass observed in nutrient-enriched marshes.
The continuous availability of high NO3
2 in the water and more
decomposable marsh grass detritus (due to higherNcontent and lower
lignin) increased decomposition rates (Table 1). Whole-ecosystem
nitrate removal was 40 times higher in the nutrient-enriched marsh
and was primarily attributable to microbial use of the added NO3
decompose organic matter12. Potential denitrification—an indicator of
anaerobic microbial decomposition using nitrate as an electron
acceptor with the end product being N2 gas—increased 1.7-fold in
creek bank sediments, while litter respiration—a measure of aerobic
microbial decomposition—almost doubled (1.9-fold). Denitrification
is the highest energy-yielding decomposition process in anoxic marsh
sediments and is favoured in the presence of high nitrate5. Accelerated
decomposition increased the fraction of fine detrital organic matter,
with 65% of the cores from nutrient-enriched creeks having a higher
percentage of fine organic matter. As a result, the fine-grained, lessconsolidated
creek banks retained more water at low tide (Fig. 2d).
The combination of fewer roots and rhizomes, drag by tidal currents
on lodged plants, more decomposed organic matter and higher water
content undermines the structural integrity of the creek bank such that
the effects of standard physical forces become enhanced. Loss of roots
1The Ecosystems Center, Marine Biological Laboratory, 7 MBL Street, Woods Hole, Massachusetts 02543, USA. 2Department of Biology, Sewanee University of the South, 735 University Avenue, Sewanee,
Tennessee 37383, USA. 3Department of Botany, Connecticut College, 270 Mohegan Avenue, New London, Connecticut 06320, USA. 4Department of Biological Sciences, Louisiana State University, Baton
Rouge, Louisiana 70803, USA. 5Department of Earth and Environment, Boston University, 675 Commonwealth Avenue, Boston, Massachusetts 02215, USA. 6Department of Natural Resources and the
Environment, University of New Hampshire, 8 College Road, Durham, New Hampshire 03824, USA.
388 | NATURE | VOL 490 | 18 OCTOBER 2012
©2012 Macmillan Publishers Limited. All rights reserved
and rhizomes (which bind sediments and provide drainage macropores)
and loss of large organic matter particles (which form air pockets
that can help drain creek banks) both contribute to increased creekbank
water content18. The higher pore water pressure in the bank
reduces the frictional shear strength of the soil and increases the sliding
force by adding weight to the creek bank19. At low water, the weight of
the saturated bank exceeds the cohesive forces holding it together, the
top of the bank cracks and creek-bank sections slide downward by
gravitational slumping18. The structural failure of the creek edge
implies that tidal forces, which under non-eutrophic conditions can
be withstood, overcome the lowered cohesive strength of the nutrientenriched
bank habitat. Cracks developed over time with nutrient
enrichment (Fig. 2d), and after seven years of enrichment, there were
more (threefold) and longer (4.5-fold) fractures at the top of the bank
parallel to the creek (Table 1) and large blocks of low marsh slumped
into the creek (Fig. 2e). Without the buttressing edge of low marsh,
high-marsh turf sheared from the sediment at the base of the active
rooting layer creating ‘toupees’ that slid down slope into creek channels
(Supplementary Fig. 1), tripling the area of bare mud over time
(Table 1). The average width of the band of tall S. alterniflora along
the creek edge decreased from approximately 3m to approximately
2m wide and became highly reticulated (Fig. 1c versus Fig. 1f; Fig. 2e)
in nutrient-enriched systems. High-resolution measurements of
channel cross-sections between the fifth and eighth years of enrichment
indicate that the maximum erosion rate of creek banks was
0.2m3 per year per metre of channel length. Loss of marsh along creek
channel edges contrasts sharply with models that suggest that higher
above-ground plant biomass in response to nutrients would in turn
trap more sediment and stabilize marsh edges relative to sea-level
rise20. Slumping and cracking leads to a positive feedback, with
increased infiltration by nutrient-rich water into sediments, which
stimulate microbial decomposition of peat and further weakens sediments.
As nutrient-enriched creek banks collapse and retreat, channels
widen, increasing the unvegetated intertidal area at the expense of
The generality of our whole-ecosystem experiment is supported by
extensive process work in small marsh fertilization plots and in anthropogenically
nutrient-enriched estuaries. From Louisiana to Nova
Scotia, nutrient enrichment has been shown consistently to increase
above-ground plant biomass14,16, mainly to decrease but sometimes
not to change below-ground biomass16,21–24 and to increase decomposition22,25.
The decomposition response might be expected to be
stronger inNorthernmarshes with a high percentage of organic matter,
but because the organic content of our creek-bank marsh is in the
middle of the range for Atlantic coast marshes (Methods), we expect
our decomposition response is typical. The combination of root loss
and increased decomposition can decrease soil strength26. The development
of cracks that lead to marsh loss has been approximately linear
over the initial nine-years; however, this is a process that will play out
over decades. Much uncertainty remains about whether this process
of marsh loss is self-limiting, because creek banks may eventually
be stabilized by the decrease in slope steepness from slumps, or selfreinforcing
owing to fractures facilitating the seepage of nutrientenriched
water into the bank and stimulating decomposition. Other
local environmental factors (such as tidal range, temperature and sediment
deposition) may modify the effects of nutrients on ecosystem
processes. Therefore, understanding the whole-ecosystem response to
nutrient loading across broad environmental gradients and longer
timescales requires more experimental ecosystem-level studies.
Many salt marshes may be at risk, because nutrient fluxes to the
coast have increased worldwide, with the largest increases in N flux
occurring at coastlines with large areas of intertidal marshland in the
temperate zones of eastern North America, Europe and eastern China
(Fig. 3a). A recent survey27 documented 415 eutrophic coastal systems
of concern worldwide, and found that only 13 systems were in recovery.
There is evidence that salt-marsh loss in Europe5 and along the Atlantic
coast of the United States 6,7 may be driven, in part, by anthropogenic
nutrient enrichment. Along Long Island Sound, Connecticut, USA,
coincident with increased total N in runoff, several marshes lost 27%–
54% of their low marsh (Fig. 3b), but very little high marsh, resulting
ratio = 1.1
ratio = 1.5
d 2010 e 2010
Figure 1 | Comparison photos of the marshes from the ecosystem nutrient-enrichment experiment. a–c, Reference. d–f, Nutrient-enriched. Photo credits:
a, b, d and e, L.A.D.; c and f, Google Earth (19 June 2010 image, copyright 2012 Google).
1 8 O C T O B E R 2 0 1 2 | VO L 4 9 0 | N AT U R E | 3 8 9
©2012 Macmillan Publishers Limited. All rights reserved
1 2 3
0 20 40
Below-ground live biomass
(g m–2 cm–1)
Years of nutrient enrichment
Fracture density in bank
Number of slumps in channel
5 6 7
0 1 2 3
Years of nutrient enrichment
53 55 58 60
Large Small Total
a b c
d Creek-bank water content (%) e f
Figure 2 | Ecosystem attributes of reference and nutrient-enriched salt
marshes after up to 7 years of nutrient enrichment. a–c, Plant attributes.
a, Annual response ratio (nutrient-enriched/reference) for above-ground shoot
attributes. Responses were comparable to other marsh nutrient-enrichment
studies along the North American Atlantic coast and, as is typical in marshes,
the strength was variable from year to year (green indicates average values of
fertilized marshes greater than reference marshes, n56 per year; solid green
line is shoot specific mass; dashed green line is shoot height; points above the
response-ratio of 1.0 (horizontal black line) indicate increased growth in
response to fertilization in that year). b, Below-ground root and rhizome
biomass in reference (open circles) and nutrient-enriched (solid green squares)
marshes (2010; n520 per treatment). c, The above-ground to below-ground
plant biomass ratio (n520 per treatment). d–f, Creek-bank geotechnical and
geomorphic attributes. d, Vertical profile of percentage water content in
reference (open circles) and nutrient-enriched (solid green squares) marsh
creek-bank sediments with depth (2010; n520 per treatment). e, Fracture
density in high marsh plotted against years of nutrient enrichment. Annual
means are shown for reference (open squares), the start of enrichment in 2009
(solid squares) and the start of enrichment in 2004 (solid circles). f, Number of
low-marsh creek-bank slumps in the channel in reference (open bars) and
nutrient-enriched (solid green bars) marshes (2010; n52). Values are all
Table 1 | Response of salt marsh ecosystem properties to chronic nutrient enrichment
Ecosystem properties Response ratio Mean (standard error) P-value
Shoot height (cm)* 1.1 146 (3) 154 (2) 0.04
Shoot mass (g)* 1.2 6.18 (0.28) 7.18 (0.18) 0.01
Shoot specific mass (g cm21)* 1.1 0.042 (0.002) 0.046 (0.001) 0.02
Lodging (%) 0 (0) 41 (2) ,0.001
Foliar N (%)* 1.1 1.43 (0.08) 1.54 (0.09) ,0.001
Lignin (%)*** 0.6 30 (7.5) 17 (1.6) 0.07
Below-ground biomass (gm22) 0.7 579 (60) 387 (64) 0.08
Creek geomorphology and sediment geotechnical properties
Fracture density (number per 50m of creek edge)** 3 1.1 (0.2) 3.3 (0.7) ,0.001
Amount of creek bank with fractures (%)** 4.5 6.6 (0.7) 29.5 (2.6) ,0.001
Fracture length (m)** 4.5 3.3 (0.4) 14.7 (1.7) 0.002
Exposed mud area (%)** 3 7.5 (1.7) 22.8 (4.6) 0.004
Total number of slumps in channel per creek 2.1 19.0 (1.0) 40.5 (11.5) 0.07
Fine organic matter (%) 1.7 16.1 (1.9) 26.7 (4.5) 0.17
Water content (%) 1.04 56.2 (0.4) 58.5 (0.1) 0.02
Channel width/depth ratio 1.3 3.02 (0.11) 3.86 (0.02) ,0.001
Microbial decomposition processes
Plant litter respiration
(CO2 g21 S21)
1.9 1.29 (0.12) 2.49 (0.33) 0.04
Potential denitrification in creekbank sediment
(nmol g21 h21)***
1.7 40.7 (7.9) 70.8 (6.3) 0.01
A response ratio (nutrient-enrichment/reference) greater than one indicates a positive response to nutrient enrichment. P-values indicate the effect of nutrient enrichment on response variables; see ‘Statistical
summary’ in Methods. Means (6s.e.m., standard error of the mean) were calculated from the data averaged by creek each year (N52 for nutrient-enriched and N52 for reference, except where noted). *Means
(6s.e.m.) calculated from data averaged by creek and pooled across seven years of data. **Means (6s.e.m.) calculated from long-term nutrient-enriched and reference data averaged by creek and pooled across
two years (2009 and 2010) of data. ***Data averaged across subplots within creeks (N51 per treatment).
390 | NATURE | VOL 490 | 18 OCTOBER 2012
©2012 Macmillan Publishers Limited. All rights reserved
in wider creeks and increased mudflat area over the last three decades7.
In Jamaica Bay, New York, USA, the rate of creek and open-bay-edge
marsh loss in the 1990s was almost double that of the previous four
decades (134km2 yr21 versus 72km2 yr21) and the timing corresponds
to an increase in total nitrogen loading (up to 80 mM open-bay total
dissolved N) from sewage inputs6. Jamaica Bay marshes also have
reduced root and rhizome mass and more degraded sediment24, similar
to our experimentally nutrient-enriched marshes.
The loss of creek-channel marsh banks is particularly significant
because it is the most productive marsh zone and a critical area for
regulating (C storage and N removal) and provisioning (fisheries) ecosystem
services. Below-ground plant productivity disproportionately
contributes to detrital organic matter in a wide variety of ecosystems28,
suggesting that the measured reduction in below-ground allocation
may decrease the C storage potential of marshes. Marshes along creek
channel edges are hotspots for denitrification with rates of N removal
fivefold to tenfold higher than mudflats29, suggesting that marsh conversion
to mudflat decreases the ability of the coastal landscape to
remove N from tidal waters. Considerable research shows the importance
of the creek marsh edges in providing shelter and food to juvenile
fish and shrimp30,31, so the loss of marsh-edge habitat may ultimately
negatively affect fisheries production.
Simultaneous increases in nutrient loading and in sea-level rise may
result in synergistic marsh loss greater than the effects of either stressor
alone. For example, the higher wave energy and flow velocities associated
with sea level rise32 when combined with decreased creek-bank
stability induced by nutrient enrichment may accelerate erosion and
creek-bank loss. The drowning of high marsh due to sea-level rise and
loss of creek-edge marsh due to eutrophication, especially when exacerbated
by upland development that limits the ability of marshes to
move inland,may lead to a coastal landscapewith a dramatically reduced
capacity to provide important ecological and economic services.Nflux
to the coastal zone has already increased at least tenfold over preindustrial
levels and is projected to continue to increase as we ramp
up fertilizer use to produce food for the expanding human population
over the next few decades1,27. The potential deterioration of coastal
marshes due to eutrophication adds an unanticipated dimension to
the challenge of managing nitrogen while meeting food production in
the twenty-first century.
We enriched primary tidal creeks in Plum Island Estuary, Massachusetts, USA, to
nutrient levels corresponding to moderately-to-highly eutrophic coastal waters by
addingNand P to the twice-daily flooding tides for nine years (2004–2012) during
the growing season (about 120 days, 15 May–15 September), enriching about
30,000m2 of marsh per experimental primary creek system (N52 enrichment
started in 2004, N51 started in 2009, reference N52–6)13. Initial measurements
(1998–2003) found few differencesamong tidal creeks13, and other potential drivers
did not differ among treatmentmarshes or do not occur in the PlumIsland Estuary
To detect changes in plant biomass allocation, the height, dry weight, and
quality (percentage N, lignin content) of Spartina alterniflora above-ground
shoots were measured and below-ground cores were analysed for live roots and
rhizome biomass. Cores were also analysed for sediment geotechnical properties
(water content, percentage organic matter and particle size). To determine
changes in creek geomorphology, fractures in the vegetated marsh platform were
enumerated along 250–300m of creek banks and point-intercept transects indicated
the presence or absence of vegetation in the creek bank. Creek-bank blocks
that had slumped into tidal creeks were enumerated and creek width, depth and
erosion measured over time using a ‘total station’ and high-precision GPS surveys.
Microbial decomposition was measured as potential denitrification in the creek
bank and microbial respiration of surface litter.
Full Methods and any associated references are available in the online version of
Received 16 April; accepted 20 August 2012.
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fluxes from pre-industrial to contemporary period
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Figure 3 | The global relationship between nutrient loading and salt-marsh
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fromcontinents to coastal oceans fromthe pre-industrial period (1800s) to the
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b, Conversion of low marsh to mudflat in six nutrient-enriched Long Island
Sound estuaries10 (see box in a indicating location of b). Substantial loss of low
marsh (solid line), smaller loss ofhighmarsh (dotted line) and increase inmudflat
(dashed line) area over time correlates with increased nutrients from sewage
treatment plants and runoff from land. Values are mean 6 standard error.
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Supplementary Information is available in the online version of the paper.
Acknowledgements We thank our TIDE (Trophic cascades and Interacting control
processes in a Detritus-based Ecosystem), LENS (Landscape Evolution in a Nutrient
enriched Saltmarsh) and PIE-LTER (Plum Island Ecosystems Long-term Ecological
Research) colleagues for field assistance and comments. We thank the many research
assistants, graduate and undergraduate students who maintained the nutrient
enrichment and analysed samples. This work is supported by grants from the NSF
(DEB0816963, DEB0213767, OCE0923689, OCE 0423565, OCE0924287), the
NOAA and The Mellon Foundation.
Author Contributions L.A.D., D.S.J., R.S.W., B.J.P. & J.W.F. designed the experiment and
participated in sampling and data analysis. S.F. participated in geomorphic and
geotechnical evaluation. W.M.W. estimated global N loading to coastal saltmarshes.
L.A.D. and D.S.J. wrote the initial manuscript. All authors contributed to and approved
Author Information The data reported in this paper are archived in the Plum Island
PIE-LTER database. Reprints and permissions information is available at
www.nature.com/reprints. The authors declare no competing financial interests.
Readers are welcome to comment on the online version of the paper. Correspondence
and requests for materials should be addressed to L.A.D. (firstname.lastname@example.org).
392 | NATURE | VOL 490 | 18 OCTOBER 2012
©2012 Macmillan Publishers Limited. All rights reserved
Site description. Our study was conducted in primary tidal creeks12,13,33,34 in the
Plum Island Estuary in northeastern Massachusetts (42u 459 N; 70u529 W). The
Plum Island Estuary is a salt-marsh-dominated systemthat is currently unaffected
by nutrient enrichment13. The system has twice-daily tides (mean tide range 2.9 m;
20–33 p.s.u.). Of the total estuarine area of 59.8km2, approximately 39.8km2
is vegetated wetlands, most of which is classic Spartina salt marsh35,36. Spartina
alterniflora (tall-form smooth cordgrass: 130–200cm in height, approximately
1,100 gm22 yr21 above-ground production) forms a twice-daily flooded 2–3-mwide
low marsh band along tidal creek channels. Spartina patens (saltmeadow
cordgrass: 20–50cm in height, approximately 600 gm22 yr21 above-ground
production) dominates the high marsh platform and is flooded by around 25%
of high tides. On creek banks and bayfronts S. alterniflora occupies the top half of
the mean tide range. S. alterniflora stems trap inorganic sediment, building steep
creek channel banks of cohesive sediment (around 15% organic matter, with the
mineral component composed of 58% silt, 25% clay and 16% sand), while belowground
roots and rhizomes (80% of root biomass in the top 20 cm, with some
rhizomes over 1m in depth) provide fibrous material that binds sediment and
stabilizes the marsh edge37–39. The percentage of organic C in the sediments of our
creek-bank marshes (7% organic C; 15% organic matter) is in the middle of
the range for salt marshes along the Atlantic and Gulf coasts of North America
(,1–28% organic C, with most marshes in the 6–15% range)40. S. patens contributes
to marsh elevation principally by below-ground biomass (80% of roots and
rhizomes are in the top 10 cm) accumulation in the form of peat.
Nutrient-enrichment experiment. Our experiment differs from others in four
important ways. (1) Nitrogen was added as nitrate (NO3
2), the form that dominates
land-derived N, contributing to coastal eutrophication and that is used
directly as an electron acceptor in decomposition processes. (2) Nutrients were
added directly to flooding tidal water to mimic the most important way in which
anthropogenic nutrient loading is delivered to tidal marshes. Dissolved nutrients
carried in water will interact and reach parts of the ecosystem differently fromdry
fertilizer. Previous plot-level salt-marsh nutrient-enrichment studies41 used dry
fertilizer (typically urea or NH4NO3) sprinkled approximately bi-weekly to
monthly on the surface of small plots (typically ,5m2) and were generally not
conducted in tall, creek-bank S. alterniflora environments. Tidal water is the primary
vector forNdelivery to coastal marshes, suggesting that if mode (dissolved in
water versus dry surface application) and type of N (NO3
important in determining ecosystem response to nutrient enrichment, previous
experiments may not be sufficient for determining how salt marshes respond to
coastal nutrient enrichment. (3) We conducted a long-term (nutrient enrichment
began in 2004) ecosystem manipulation experiment in which marsh landscapes
were nutrient-enriched to levels corresponding to moderate to highly eutrophic
coastal waters42–44. Our NO3
2 enrichment target of 70–100 mMNO3
2 (added as
NaNO3) was 15 times the Plum Island background (,5 mMNO3
2) and the PO4
(added as NaH2PO4) target of 5–7 mM was 5 times the background (,1 mM
32). This was approximately a 15:1 N:P ratio in flooding creek water. (4)
We conducted an ecosystem-level experiment consisting of experimental marsh
units (n56; 3 reference and 3 nutrient-enriched) comprised of first-order creeks
(about 300mlong and 15mwide at the mouth, tapering to 2mnear terminus) and
about 30,000m2 of cordgrass marsh area, thus allowing us to examine interacting
habitats in the marsh ecosystem (creek channels, mudflats, creek-bank low marsh,
and high marsh) and examine the response of plants, animals, biogeochemical
processes and landscape-level geomorphic processes. Other factors did not
differ among treatment creeks (Supplementary Information). Analysis of baseline
characteristics before experimentalmanipulation (1998–2003) found little difference
among the experimental marsh systems13. The primary comparisons are among
two long-term (7 years by the end of 2010) nutrient-enriched (N1, N2) marshes
and two reference (R1, R2) marshes that have been intensively monitored.
To provide a time series of geomorphic change, we include data from a third
nutrient-enrichment marsh (N3) that was started in 2009.
Measurements. This is a multi-year nutrient enrichment experiment, and not all
response variables reported here were measured in each year. Some responses to
enrichment were unanticipated (for example, rapid geomorphic changes) and so
measurements were not taken before manipulation and sometimes only a single
season of data is available. The nature of the different data sets with different time
series of collections necessitated various statistical analyses (detailed below).
Above-ground plant responses. Above-ground measurements of creek-bank
S. alterniflora were taken in creeks R1, R2 and N1, N2 in years 1–7 of nutrient
enrichment. Individual plant shoot length (cm), shoot mass (g dry weight) and
shoot-specific mass (g dry weight cm21) of creek-bank S. alterniflora were measured
at the end of the growing season (mid-August) on individual shoots (18–25)
at three sites within each creek (n554–75 shoots per treatment creek per year).
Each shoot was individually washed to remove sediments, measured for length,
dried at 80 uC to a constant mass and weighed. Leaf tissue from 3–5 leaves from
each site was ground and analysed for percentage nitrogen using a PerkinElmer
2400 Series II CHNS/O analyser (n54–8 per creek per year).
In year 5 of nutrient enrichment, lignin (as a percentage of the ash-free organic
content) was determined on composite samples of 3–5 shoots from three sites
within creeks R1 and N1 as acid-insoluble fractions using a two-stage digestion in
Plant lodging was surveyed at the end of the season after peak production in N1,
N2, R1, R2 and four additional reference creeks in nutrient treatment year 4 (N52
for nutrient enrichment and N56 for reference). Surveys were completed on
10-m sections every 50m from the 0-m mark to 300m landward. Each section
classified into a lodging class (0–5, 5–15, 25–50 and .50% of plants in the area
lodged) for a sampling effort of n516–26 sections per creek.
Below-ground responses. Below-ground biomass, organic matter and water
content were determined by coring (n510 per creek, 10cm diameter, taken to
a depth of about 0.5 m; creeks R1, R2, N1, N2) in treatment year 7 (2010). Cores
were sliced into sections (0–5, 5–10, 10–20, 20–30, 30–40 cm), sub-samples were
taken for determination of percentage water (a small syringe core in each section),
and the remaining material was separated by sieving into two size classes of dead
organic matter (large .3mm; fine ,3mm and .1mm) and live below-ground
biomass (roots and rhizomes). Sediment geotechnical properties (percentage
water was determined as mass loss after drying a known volume of sediment at
105 uC for 24 h; percentage fine organic matter was taken to be detritus greater
than 1mmbut less than 3mmin size) were determined on cores taken for belowground
plant biomass. For statistical analysis (see ‘Statistical Summary’ below), we
focused on the top 20 cm of the cores.
Microbial decomposition processes. Total microbial production in surficial sediments
was 54% higher46 (years 1 and 2) and potential denitrification on the high
S. patens marsh was higher47 (year 3) than in the reference systems. To determine
whether microbial denitrification was also increased in creek banks, potential
denitrification48 was measured on sediment slurries—at the surface (0–5 cm)
and deep (5–10 cm)—from creek-bank cores from three sites in creeks R1 and
N1 in year 5 of nutrient enrichment (n53 per depth per creek).
To determine whether plant litter decomposition was accelerated, in year 7
plant litter respirationwasmeasured.Respirationwasmeasured fromdecomposing
litter from litterbags (15 g dry weight of S. patens; 1mmmesh size) placed flush on
the high-marsh (S. patens) surface in nutrient enrichment (N1, N2) and reference
(R1, R2) creeks. Nutrient enrichment stimulated detritivore snail densities34 and
therefore to account for the effect of detritivore density on decomposition, litterbags
were manipulated to have snail densities of 0, 1, 2, 4 or 8 times the reference
creek densities (n55–8 litterbags per creek). After 5 weeks, microbial respiration
(CO2 g21 s21) of 2–3 g of litter from the litterbags was measured using a LI-6200
Portable Photosynthesis System.
Creek-bank fracture density and vegetation loss. These measurements were
taken during the growing season in creeks R1, R2 and N1, N2 in years 6 and 7
of nutrient enrichment. Fractures, defined as a visible break in the high marsh
(S. patens-dominated) turf that parallels the creek channel (Supplementary Fig. 1)
within 3m of the S. alterniflora/S. patens border were measured early in the
growing season before cordgrass growth obscured these features. Both sides of
each creek were sampled for fractures in contiguous 50-msegments from the 0-m
mark to 200-m landward. The number of fractures and their characteristics
(length, width and depth of fracture) were recorded within each segment. In these
same segments, percentage exposed sediment (mud) area was determined by
point-intercept transects in the middle of the growing season, when the grass
canopy was fully developed, at 1-mincrements. The soil surface 1-m perpendicular
and creekward of the S. alterniflora/S. patens border was scored as ‘vegetated’ (with
S. alterniflora culms within a 30-cm-diameter circle of the point) or as ‘bare mud’
(without S. alterniflora). The fraction of points unvegetated within each
50-m segment was considered a single observation. For fracture density, fracture
length, percentage fractured and percentage exposed mud, each 50-m segment
(4 per channel side, 2 sides) was considered an individual observation, thus providing
n58 observations per creek per year. This was also done for the third
enriched and reference creeks (R3, N3) in years 1–4 of enrichment.
Creek-bank and channel structure. High-resolution total station surveys of
reference (R1, R2) and nutrient-enriched creek channels (N1, N2) and banks were
performed in years 5 (2008) and 8 (2011) of the nutrient-enrichment experiment.
We initially measured 43 cross-sections in enriched creeks and 38 in the reference
channels, with 10–20 points per cross-section, depending on the morphological
complexity of the cross-section. Twenty-six cross-sections in the nutrientenriched
creek channels and 27 in the reference channels were reoccupied after
three years, whereas in the remaining cross-sections the control poles were lost.
Erosion was computed for each cross-section as the areal difference between the
year 5 and year 8 surveys.
©2012 Macmillan Publishers Limited. All rights reserved
In year 7 of nutrient enrichment slumped sections of creek bankwere enumerated
in creek channels (R1, R2 and N1, N2). At low tide, large and small slumped
sections of creek bank were enumerated in 50-m reaches starting from the 0-m
mark up to 200–250 m. Large sections were defined as peat blocks that were
separated away from the low marsh area by at least 0.25m horizontally and at
least 0.25mlower than themean elevation of the lowmarsh. Large peat blocks were
at least 1meither in height, length or width, with small blocks being at least 0.25m
in height. Height was defined as the distance from the bottom of the peat block to
the highest point of the block. Ninety per cent of the large peat blocks had live
S. alterniflora shoots and were 1.1m wide and 2.1m long on average. Small creekbank
sections were defined as low-marsh peat chunks that were ,1m in height,
width and length, but at least 0.25m in at least one of these dimensions. Small
slumps were generally unvegetated, had visibly eroded perimeters and were
found in the deepest part of the channel. The total count of slumps per creek
was considered a sample and thus within treatment area sampling replication
was n51 per creek.
Global nutrient loading to salt marshes. We used an existing global river network
N removal model49,50 to estimate global increase in N loading to coastal
oceans from the pre-industrial period (1800s) to the contemporary period
(2000s) compared to the locations of salt marshes5,51.
Statistical summary. Analyses were performed on data primarily from the longterm
nutrient-enrichment creeks and the paired reference creeks. Data collection
from creeks often entailed sampling several subplots within experimental creeks
(see ‘Measurements’ section above). Except where noted, data were averaged
across subplots within each creek before analysis52 and statistical analyses were
performed at the creek level (nutrient enrichment n52, reference n52). All data
were checked for assumptions of normality and homoscedasticity and transformed
tomeet assumptions53. The large spatial scale of the experiment necessitated
low replication, which can reduce statistical power, so results were considered
significant at a#0.10, as is typical in a complex large-scale ecosystem experiment
where background variability is generally high and replication low54–57. Our results,
however, are robust, because 13 out of 17 tests were significant at a#0.05 despite
low replication (Table 1). Statistical analyses were performed in R (version 2.15.0).
Repeated measures analysis of variance with between-subject factors (nutrient
enrichment) were performed for response variables for which 7 years of data were
available: plant shoot height, shoot mass, shoot specific weight (natural logX11
transformation), and percentage foliar nitrogen (arcsine-squareroot transformation).
Using ‘space for time’ substitution, linear regression models were used to
analyse data for fracture density (natural logX11 transformation), fracture
length (natural logX11 transformation), percentage of creek bank with fractures
(arcsine-squareroot transformation), and percentage non-vegetated exposed mud
area (arcsine-squareroot transformation) against the number of years the creek
had received nutrient enrichment. Data for fracture density was collected in 2009–
2012 and analysis was at the creek level, thus n524: for year 0, n53 per year (R1,
R2, R3 in 2009–2012); for years 1–4, n51 per year (N3 in 2009–2012); and for
years 6–9, n52 per year (N1, N2 in 2009–2012). Data for other variables was
collected in 2009 and 2010 and analysis was at the creek level, thus n510: for year
0, n54 (two years of data from R1, R2); for year 1, n51 (N3 in 2009); for year 2,
n51 (N3 in 2010); for year 6, n52 (N1, N2 in 2009); and for year 7, n52 (N1,
N2 in 2010).
One tailed t-tests were performed on the following response variables: percentage
water content (arcsine-squareroot transformation), live below-ground
biomass, percentage fine organic matter (arcsine-squareroot transformation),
slumps per creek (logX11 transformation), lodging (percentage of plants lodged;
arcsine-squareroot transformation), potential denitrification and percent foliar
lignin (arcsine-squareroot transformation). For lodging, t-tests were performed
on the percentage of plots within a creek that were scored with .50% plants
lodged. For potential denitrification, initial statistical analysis indicated no difference
between depths, so data from different depths were pooled in the final
analysis (n56 per creek). Because potential denitrification and percentage foliar
lignin were taken only from one creek per treatment (R1 and N1), data were
analysed with the subplots as the experimental units. These process measurements
are used as supporting evidence, but cannot be used to extrapolate the results to a
wider population of systems because there was no treatment replication at the
Analysis of covariance was performed on plant litter respiration and channel
width/depth ratio. In these analyses the factor was nutrient level—reference
(n52) and nutrient enrichment (n52). The covariate in the analysis of plant
litter respiration was the number of snails in the litterbags. The covariate in the
analysis of the channel width/depth ratio was the distance upstream from the
beginning of the treatment area (designated as 0 m; a spatial covariate). In both
analyses of covariance, the slopes were similar, but the intercepts were different,
indicating a difference between treatments at the zero covariate level.
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