carbon sequestration in wetland soils of the northern gulf of mexico coastal region

ORIGINAL PAPER

Carbon sequestration in wetland soils of the northern Gulfof Mexico coastal region

Virginia D. Hansen • Janet A. Nestlerode

Received: 5 April 2013 / Accepted: 14 November 2013

� Springer Science+Business Media Dordrecht (outside the USA) 2013

Abstract Coastal wetlands play an important but

complex role in the global carbon cycle, contributing

to the ecosystem service of greenhouse gas regulation

through carbon sequestration. Although coastal wet-

lands occupy a small percent of the total US land area,

their potential for carbon storage, especially in soils,

often exceeds that of other terrestrial ecosystems.

More than half of the coastal wetlands in the US are

located in the northern Gulf of Mexico, yet these

wetlands continue to be degraded at an alarming rate,

resulting in a significant loss of stored carbon and

reduction in capacity for carbon sequestration. We

provide estimates of surface soil carbon densities for

wetlands in the northern Gulf of Mexico coastal

region, calculated from field measurements of bulk

density and soil carbon content in the upper 10–15 cm

of soil. We combined these estimates with soil

accretion rates derived from the literature and wetland

area estimates to calculate surface soil carbon pools

and accumulation rates. Wetlands in the northern Gulf

of Mexico coastal region potentially store

34–47 Mg C ha-1 and could potentially accumulate

11,517 Gg C year-1. These estimates provide impor-

tant information that can be used to incorporate the

value of wetlands in the northern Gulf of Mexico

coastal region in future wetland management deci-

sions related to global climate change. Estimates of

carbon sequestration potential should be considered

along with estimates of other ecosystem services

provided by wetlands in the northern Gulf of Mexico

coastal region to strengthen and enhance the conser-

vation, sustainable management, and restoration of

these important natural resources.

Keywords Gulf of Mexico �Coastal wetlands �Carbon

Introduction

While recent attention has been given to the potential

for natural and agricultural soils to sequester carbon

and mitigate greenhouse gas emissions, much of this

attention has been focused on terrestrial ecosystems

(i.e., forests, grasslands, and croplands) (Jobbagy and

Jackson 2000; Post et al. 2004; Powlson et al. 2011;

USEPA 2011). Wetlands also play an important but

complex role in the global carbon cycle, contributing

to the ecosystem service of greenhouse gas regulation

through carbon sequestration (Bridgham et al. 2006;

Chmura et al. 2003; DeLaune and White 2012; Engle

2011; Mcleod et al. 2011). Wetlands may serve as

carbon sinks because they store large amounts of

carbon in aboveground biomass (e.g., forested wet-

lands) and soils (e.g., peatlands, coastal marshes);

however, wetlands can also emit significant quantities

V. D. Hansen (&) � J. A. Nestlerode

Gulf Ecology Division, US Environmental Protection

Agency, 1 Sabine Island Dr., Gulf Breeze, FL 32561,

USA

e-mail: [email protected]

123

Wetlands Ecol Manage

DOI 10.1007/s11273-013-9330-6

of methane (CH4), and nitrous oxide (N2O) to the

atmosphere (Bridgham et al. 2006; Mitsch et al. 2012).

Wetlands that occur in the northern Gulf of Mexico

coastal region include salt marshes, bottomland hard-

wood swamps, fresh marshes, mangroves, and other

types of emergent and forested wetlands (Stedman and

Dahl, 2008). Coastal wetlands gradually accrete

sediment and organic matter and have unique biogeo-

chemical characteristics which enhance their potential

for carbon storage, often exceeding that of other

terrestrial ecosystems (Bridgham et al. 2006; Call-

away et al. 1997; Choi and Wang 2004; Craft et al.

1993; Mcleod et al. 2011; Rabenhorst 1995; Whiting

and Chanton 2001). Tidal, saline wetlands, in partic-

ular, continuously accrete and bury sediments that are

rich in organic carbon while emitting negligible

amounts of greenhouse gases such as CH4 because

of the saline, anaerobic environment (Poffenbarger

et al. 2011; Whiting and Chanton 2001). In contrast,

accretion of carbon in the organic soils of coastal

freshwater wetlands (e.g., fresh marshes and forested

swamps) often occurs at slower rates and with higher

CH4 release than tidal, saline wetlands (Nyman et al.

2006; Yu et al. 2006).

More than half of the coastal wetlands in the US are

located in the northern Gulf of Mexico region (Field

et al. 1991), yet these wetlands continue to be lost at an

alarming rate (*25,000 ha year-1; Stedman and Dahl

2008) because of land use changes, coastal develop-

ment, hurricanes, sea-level rise, and subsidence (Day

et al. 2000; DeLaune and White 2012; Turner 1997).

The loss of wetlands in the northern Gulf of Mexico

coastal region results in a significant loss of stored

carbon, and further degradation of these wetlands

reduces their capacity for carbon sequestration (DeLa-

une and White 2012; Engle 2011). In order to estimate

the magnitude of carbon storage in wetlands in the

northern Gulf of Mexico coastal region, however,

quantitative estimates of soil carbon pools and carbon

sequestration rates are needed. Engle (2011) provided

gross estimates of soil carbon pools and accumulation

rates for Gulf of Mexico salt marsh and mangrove

wetlands by applying average stock and rate estimates

from the literature to wetland area estimates. The

average soil carbon pools and average soil carbon

accumulation rates for wetlands in the northern

Gulf of Mexico coastal region were estimated to be

275 Mg C ha-1 and 2.6 Mg C ha-1 year-1 in salt

marshes and 203 Mg C ha-1 and 2.1 Mg C ha-1 year-1

in mangroves (Engle 2011; Mg = 106 g). The loss of

18,385 ha of estuarine emergent and shrub wetlands in

the Gulf of Mexico from 1998 to 2004 (Stedman and

Dahl 2008), would have resulted in a 47 Gg C reduction

in soil carbon accumulation (Gg = 109 g; Engle 2011).

Engle (2011) recognized, however, that these estimates

of soil carbon pools and accumulation rates have large

uncertainties because of the paucity of quantitative data

for wetlands in the northern Gulf of Mexico coastal

region.

The purpose of this study was to provide more

refined estimates of soil carbon data for wetlands in the

northern Gulf of Mexico coastal region to improve

current carbon sequestration estimates. The rate of

carbon accumulation in wetland soils is a function of

soil carbon densities and rates of vertical soil accre-

tion. Chmura et al. (2003) presented soil organic

carbon density and accumulation rate estimates for

tidal saline wetland soils in the Gulf of Mexico;

however, most of these estimates were calculated from

reported measurements of organic matter density that

were converted to organic carbon density using the

formula from Craft et al. (1991) for salt marshes or

multiplying by a factor of 1.724 (Allen 1974) for

mangroves.

In this study, we provide estimates of soil carbon

density based on actual measurements of bulk density

and carbon content in wetland soils from a range of

wetland types in the northern Gulf of Mexico coastal

region. We then combined the soil carbon density

estimates with vertical accretion rates derived from the

literature to estimate soil carbon accumulation rates.

These estimates of soil carbon sequestration provided

by wetlands in the northern Gulf of Mexico coastal

region could be used to strengthen and enhance the

conservation, sustainable management, and restora-

tion of these important natural resources.

Methods

The US Environmental Protection Agency (USEPA)

Office of Research and Development and the US

Geological Survey (USGS) National Wetlands

Research Center collaboratively conducted a regional

pilot survey of the ambient environmental condition of

coastal wetlands in the US Gulf of Mexico (Nestlerode

et al. 2009). The target population for this survey

included wetlands within coastal watersheds in the


123

northern Gulf of Mexico region from the Rio Grande,

Texas to Florida Bay, Florida. The survey design was

two-stage: the first stage was based on the National

Wetlands Inventory Status and Trends (NWI S&T)

stratified random survey design (Dahl and Bergeson

2009), and the second stage involved a random

selection of wetland sites from plots identified in the

first stage. A northern Gulf of Mexico coastal bound-

ary was defined by all USGS 8-digit hydrologic unit

code (HUC) watersheds that intersected with the US

Gulf of Mexico coastline that were modified, where

necessary, by the boundaries of the National Oceanic

and Atmospheric Administration (NOAA) Coastal

Assessment Framework Estuarine Drainage Areas

(Fig. 1). The sample frame for this survey was

comprised of 1,071 NWI S&T plots that fell com-

pletely within the northern Gulf of Mexico coastal

watersheds. Land cover within these 10.46 km2-plots

was represented by multiple polygons in a GIS

database, and each polygon was classified as one of

several estuarine, palustrine, or upland habitat types

according to the Cowardin wetlands classification

system framework (Cowardin et al. 1979). The four

major wetland classes targeted for sampling included

estuarine emergent, estuarine shrub/scrub, palustrine

emergent, and palustrine forested habitats (Cowardin

et al. 1979). An unequal probability general random

tessellation stratified (GRTS) survey design (Stevens

and Olsen 2004) was used to select 100 survey sites

(unique latitude and longitude coordinates) within

polygons that contained targeted wetland types

(Fig. 1).

The survey evaluated wetland condition using a

‘‘three-tier framework’’ assessment strategy (Fen-

nessy et al. 2004, 2007; USEPA 2006). This approach

implemented three tiers of evaluation that varied in

Fig. 1 Gulf of Mexico (GOM) coastal wetland survey sites

shown within the boundary of GOM coastal watersheds as

defined by 12-digit HUCs and NOAA estuarine drainage areas.

Symbols represent estuarine emergent (filled triangle), estuarine

shrub (filled square), palustrine emergent (open circle), and

palustrine forested (filled plus) wetland classes


123

spatial scale and laboratory/field sampling effort

(Nestlerode et al. 2009). Broad, landscape assessments

(Tier 1) used readily available GIS and remote sensing

data; Tier 2 comprised rapid field assessment methods

incorporating simple observational measures and a

large degree of professional judgment; and Tier 3

included intensive, on-site collection of detailed

biological, physical, and chemical measurements.

Each site was surveyed once during the summers of

2007 or 2008. This paper focuses on the Tier 3 metrics

that were collected to determine soil carbon pools and

to compute soil carbon accumulation rates, using

measurements of bulk density, total carbon (TC), and

total organic carbon content (TOC).

At each site, we defined an assessment area (AA) as

a fixed area delineated by a 25 m radius circle around

the latitude and longitude coordinates generated by the

probability-based survey design. Pore water and soil

measures were collected from three random 0.25-m2

sub-plots within the AA and we assumed minimal

spatial variability within and between sub-plots. One

soil core was collected from each of the three random

sub-plots within the AA and composited for TC and

TOC analysis. One additional intact core was collected

from within one of the three 0.25-m2 plots for bulk

density determination. At sites that were surveyed in

2007, all soil cores were collected to a depth of 10 cm

using 60 ml disposable syringes with the distal tips

removed from near to surface undisturbed soil. In

2008, all soil cores were collected to a depth of 15 cm

using a stainless steel split-core sampler. Soil samples

were placed in sealable plastic bags, and stored in the

dark at 4 �C until transport to laboratories for

subsequent processing within 8 weeks of collection.

All soil samples were oven-dried prior to analysis.

Soil bulk density was determined from the dry weight

to volume ratio (Blake and Hartge 1986) and soil

moisture was determined through water loss upon

drying at 90 �C for a minimum of 24 h or until a

constant weight was achieved. Soil TC was measured

using the Micro-Dumas combustion method on a

Carlo-Erba C/H/N analyzer equipped with a high

temperature induction furnace that converted all

carbon in the sample to CO2 gas which was measured

by thermal conductivity detector (Tiessen and Moir

1993). A subsample was ashed at 500 �C for 4 h to

remove organic carbon and the ashed sample was

analyzed on the C/H/N analyzer to determine soil total

inorganic carbon (Karam 1993). Soil TOC was

calculated as the difference between TC and total

inorganic carbon. Soil carbon density (g cm-3) was

calculated as the product of percent TC or TOC and

bulk density.

To estimate soil carbon accumulation rates, vertical

accretion rates were obtained from the literature. A

comparison of dominant plant species, geography, and

salinity between wetlands from our survey and those

from the literature appropriate for wetlands in the

northern Gulf of Mexico coastal region resulted in the

identification of eight wetland sub-classes for which

vertical accretion rates were available: (1) fresh

marshes dominated by Cladium sp. (e.g., Everglades);

(2) fresh marshes dominated by other vegetation

species; (3) brackish marshes dominated by Spartina

patens; (4) Louisiana (LA) salt marshes; (5) salt

marshes outside LA; (6) bottomland hardwood wet-

lands; (7) cypress–tupelo wetlands; and (8) man-

groves. Different accretion rates were assigned to our

study sites based on average accretion rates from the

literature for these eight wetland sub-classes

(Table 1). For most of the wetland sub-classes,

vertical accretion rates determined by the 137Cs

method were obtained; however, for the two palustrine

forested wetland sub-classes, only accretion rates

determined by the feldspar marker method were

available. Accretion rates from the literature for

managed, impounded, treatment, or experimentally

nutrient-enriched wetlands were not included in the

average accretion rates used for this study because

they tended to be much greater than the accretion rates

in natural wetlands, and our survey did not target

managed wetlands. Several wetland sites in our survey

had dominant vegetation that did not match any of the

eight wetland sub-classes for which accretion rates

were available from the literature; therefore, no

accretion rates were assigned to those sites, which

included brackish marshes dominated by species other

than S. patens (e.g., Echinochloa walteri or Schoen-

plectus americanus), freshwater ponds, hydric pine

forests (dominated by Pinus sp.), titi swamps (dom-

inated by Cyrilla racemiflora), and other freshwater

shrub swamps (dominated by Baccharis sp.).

Soil carbon accumulation rates (g m-2 year-1)

were calculated as the product of soil carbon density

(g cm-3) and vertical accretion rates (cm year-1)

according to the methods presented by Chmura et al.

(2003). The total soil carbon pool and annual

accumulation of carbon in Gulf of Mexico coastal


123

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123

wetland soils was determined by multiplying average

soil carbon densities and accumulation rates by the

area of coastal wetlands in the Gulf of Mexico. Area

estimates for the four original wetland classes—

palustrine emergent, estuarine emergent, palustrine

forested, and estuarine shrub—were reported for 1998

and 2004 by Stedman and Dahl (2008). We recom-

puted the average soil carbon densities and accumu-

lation rates for these four classes of wetlands and

multiplied by the 2004 wetland area estimates (Sted-

man and Dahl 2008) to obtain soil carbon pools and

total annual soil carbon accumulation estimates.

All statistical analyses were conducted in SAS�

(version 9.2; � SAS Institute, Cary, NC). The

significance of linear regressions and t tests were

determined at a = 0.05. Significant differences of

mean values among wetland classes were determined

using Duncan’s multiple range test.

Results

We successfully surveyed and collected soil samples at

88 of the 100 base sites in the original survey design as

well as eight reference sites. Average bulk density, soil

TC and TOC fractions, and soil carbon (C) and organic

carbon (OC) densities are reported for all wetland sub-

classes (Table 2). Percent TOC was only slightly less

than percent TC in all samples (paired t test; mean

difference = 0.6271; p = 0.0043) indicating that the

inorganic carbon content is very low in most of these

wetland soils (Fig. 2a). Percent TOC was inversely

related to ln(bulk density) (TOC: p \ 0.0001,

R2 = 0.70; Fig 2b). The sites that did not show this

relationship (i.e., outliers in Fig. 2a) were located in

fresh marshes dominated by Cladium and in man-

groves in Florida; these sites had much higher TC than

TOC, reflecting higher inorganic carbon content.

Organic soils generally had low bulk density and high

TOC content while mineral soils had high bulk density

and low TOC content (Fig. 2b). Average soil OC

densities ranged from 0.021 to 0.090 g OC cm-3

(Table 2) but were not significantly different among

the wetland sub-classes (Fig. 3). Variability in bulk

density, TC and TOC content and C and OC densities

existed within wetland sub-classes, however.

Three of our sites were located in fresh marsh

dominated by Cladium; one in Big Cypress National

Table 2 Average bulk density, TC and TOC content and densities in Gulf of Mexico coastal wetland soils (mean ± 1SD)

Wetland

class

Wetland sub-class Number

of sites

Bulk density

(g cm-3)

Total carbon Total organic carbon

TC (%) Density

(g C cm-3)

TOC (%) Density

(g OC cm-3)

Palustrine

emergent

Fresh marsh dom. by

Cladium

3 0.68 ± 0.61 11.36 ± 8.01 0.045 ± 0.017 4.16 ± 1.91 0.021 ± 0.007

Fresh marsh dom. by

other

13 0.57 ± 0.55 20.33 ± 18.64 0.031 ± 0.023 20.07 ± 18.55 0.030 ± 0.024

Freshwater pond 1 1.11 3.44 0.038 3.42 0.038

Estuarine

emergent

Brackish marsh dom.

by Spartina patens

16 0.52 ± 0.42 13.37 ± 10.37 0.040 ± 0.024 13.32 ± 10.34 0.040 ± 0.024

Brackish marsh dom.

by other

3 0.02 ± 0.09 20.88 ± 12.18 0.024 ± 0.004 20.80 ± 12.15 0.024 ± 0.004

Salt marsh in LA 8 0.40 ± 0.26 9.33 ± 4.54 0.035 ± 0.023 9.31 ± 4.53 0.035 ± 0.023

Salt marsh outside LA 7 0.56 ± 0.36 8.15 ± 6.54 0.031 ± 0.023 8.10 ± 6.51 0.030 ± 0.023

Palustrine

forested

Freshwater shrub 2 0.79 ± 0.26 10.04 ± 6.41 0.071 ± 0.021 9.92 ± 1.68 0.070 ± 0.021

Titi swamp 5 0.79 ± 0.26 14.67 ± 14.83 0.092 ± 0.080 14.36 ± 14.61 0.090 ± 0.079

Hydric pine 3 1.15 ± 0.24 2.04 ± 1.68 0.026 ± 0.021 2.03 ± 1.68 0.026 ± 0.021

Bottomland hardwood 13 0.78 ± 0.30 8.48 ± 9.00 0.048 ± 0.032 8.36 ± 8.91 0.047 ± 0.032

Cypress–tupelo 4 0.14 ± 0.04 35.58 ± 13.53 0.045 ± 0.012 34.69 ± 13.04 0.044 ± 0.012

Estuarine

forested

Mangrove 10 0.22 ± 0.26 23.94 ± 7.86 0.041 ± 0.023 21.77 ± 10.75 0.030 ± 0.006


123

Preserve, Florida and two in Everglades National Park,

Florida. The site in Big Cypress National Preserve

showed no difference between soil C (0.029 g C cm-3)

and OC (0.028 g OC cm-3) density. At the sites in

Everglades National Park, however, soil C density

(0.042–0.063 g C cm-3) was much higher than soil OC

density (0.016–0.018 g OC cm-3) reflecting the high

percent of inorganic carbon in these soils.

Fresh marshes dominated by plant genera other than

Cladium (e.g., Panicum, Phragmites, Eleocharis, Leer-

sia, and Sagittaria) showed no significant differences

between soil C and OC density. In general, sites in

Louisiana and Texas had lower OC densities

(0.006–0.021 g OC cm-3) than sites in Florida

(0.013–0.095 g OC cm-3). The site with the lowest

OC density (0.006 g OC cm-3) was dominated by the

invasive plant, Phragmites australis.

Brackish marshes were primarily dominated by S.

patens and were only sampled in Louisiana and Texas.

There were no significant differences between soil C

and OC density in brackish marshes or salt marshes.

Soil OC densities did not differ between salt marshes in

Louisiana (0.015–0.078 g OC cm-3) and salt marshes

outside of Louisiana (0.009–0.078 g OC cm-3) or

between dominant plant communities: Spartina alter-

niflora (0.015–0.078 g OC cm-3), Juncus roemeri-

anus (0.019–0.078 g OC cm-3).

Cypress–tupelo swamps were dominated by Taxo-

dium distichum and Nyssa aquatica while bottomland

hardwood swamps were typically dominated by Acer

Fig. 2 Relationships between a TC % and TOC %) (p \ 0.0001; R2 = 0.97) and b ln bulk density (g cm-3) and TOC % (p \ 0.0001;

R2 = 0.70), in Gulf of Mexico coastal wetland soils. Symbols represent organic soil (open square) and mineral soil (open triangle)

Fig. 3 Box plots of soil

organic carbon density in

Gulf of Mexico coastal

wetlands (diamond mean;

line median; boxes 25th and

75th percentiles; whiskers

minimum and maximum)


123

rubrum, Fraxinus sp. or Quercus sp. (Table 1).

Although mean soil OC density was not significantly

different between these forested wetlands, cypress–

tupelo swamp soils had a smaller range of soil OC

density (0.029–0.053 g OC cm-3) than bottomland

hardwood soils (0.013–0.104 g OC cm-3). The low-

est soil OC density (0.013 g OC cm-3) for forested

wetlands was found at a degraded wetland dominated

by the invasive species, Triadica sebifera (Chinese

tallow).

Mangrove sites were all located in Florida and

were primarily dominated by Rhizophora mangle.

In general, soil C density did not differ from soil

OC density at these sites. However, one site in

southern Everglades National Park, that was dom-

inated by Laguncularia racemosa and Salicornia

sp., had high bulk density (0.94 g cm-3) with low

TOC content (3 %) compared to other mangrove

sites (bulk density 0.11–0.25 g cm-3 and TOC

9–37 %); soil C density (0.105 g C cm-3) was

much higher than soil OC density (0.030

g OC cm-3) at this site.

Accretion rates for eight wetland sub-classes were

identified by matching wetland type, dominant plant

species, and porewater salinity from our survey to

accretion rate studies from the literature (Table 1).

Average soil accretion rates from the literature were

lowest for bottomland hardwood forests, mangroves,

and fresh marshes dominated by Cladium sp.

(0.23–0.30 cm year-1; Table 1). Salt marshes in

Louisiana had a higher average soil accretion rate

(0.94 cm year-1) than salt marshes elsewhere in the

Gulf of Mexico (0.57 cm year-1; Table 1). Fresh

marshes that were dominated by plant species other

than Cladium sp. had an average accretion rate

(0.84 cm year-1) similar to cypress–tupelo swamps

(0.85 cm year-1; Table 1).

Within wetland sub-classes, soil carbon accumulation

rates were linearly related to soil carbon density as a

constant average accretion rate was applied to all sites

within each wetland sub-class. Average soil carbon

accumulation rates ranged from 103 to

381 g C m-2 year-1 for TC and 47–372 g OC m-2 -

year-1 for TOC (Table 3). Carbon and OC accumulation

rates were highest for salt marshes in Louisiana and

cypress–tupelo wetlands and lowest for fresh marshes

dominated by Cladium and mangrove wetlands

(Table 3). Average C accumulation rates differed from

OC accumulation rates only in fresh marshes dominated

by Cladium and mangrove wetlands.

By applying the average rates of soil C accumulation to

the reported wetland area from 2004 (Stedman and Dahl

2008), we estimate that the 5,308,468 ha of wetlands in

the northern Gulf of Mexico coastal region potentially

accumulate 11,517 Gg C year-1 (Gg = g 9 109) with

approximately 46 % contributed by palustrine and estu-

arine emergent wetlands combined (2,663 and

2,664 Gg C year-1, respectively) and 51 % contributed

by palustrine forested wetlands (5,899 Gg C year-1)

(Table 4). Because of the small area of estuarine shrub

wetlands (i.e., mangroves) in the Gulf of Mexico, this

wetland class contributed only 291 Gg C year-1.

Table 3 Average total carbon (TC) and total organic carbon (TOC) densities from Table 2, accretion rates from Table 1, and

calculated average soil C and OC accumulation rates in Gulf of Mexico coastal wetland soils (mean ± 1SD)

Wetland

class

Wetland sub-class Density (g C cm-3) Accretion rate

(cm year-1)

Accumulation rate

(g m2 year-1)

TC TOC C OC

Palustrine

emergent

Fresh marsh dom. by

Cladium

0.045 ± 0.017 0.021 ± 0.007 0.23 ± 0.06 102.8 ± 39.1 47.4 ± 15.9

Fresh marsh dom. by other 0.031 ± 0.023 0.030 ± 0.024 0.84 ± 0.19 260.3 ± 196.4 250.5 ± 199.1

Estuarine

emergent

Brackish marsh dom. by

Spartina patens

0.040 ± 0.024 0.040 ± 0.024 0.72 ± 0.30 290.8 ± 170.3 289.6 ± 169.6

Salt marsh in LA 0.035 ± 0.023 0.035 ± 0.023 0.94 ± 0.32 329.2 ± 217.1 328.3 ± 216.6

Salt marsh outside LA 0.031 ± 0.023 0.030 ± 0.023 0.57 ± 0.18 173.9 ± 131.8 173.1 ± 131.7

Palustrine

forested

Bottomland hardwood 0.048 ± 0.032 0.047 ± 0.032 0.30 ± 0.04 142.6 ± 97.3 140.5 ± 96.5

Cypress–tupelo 0.045 ± 0.012 0.044 ± 0.012 0.85 ± 0.54 380.9 ± 103.5 371.7 ± 98.3

Estuarine

forested

Mangrove 0.041 ± 0.023 0.030 ± 0.006 0.26 ± 0.11 105.7 ± 60.1 78.9 ± 16.3


123

We also estimated the amount of C currently stored

in surface soils of wetlands in the northern Gulf of

Mexico coastal region from average soil C densities

for each wetland class and the area of coastal wetlands

(Stedman and Dahl 2008). Using the 2004 estimates

for wetland area from Stedman and Dahl (2008),

wetlands in the northern Gulf of Mexico coastal region

store 34–47 Mg C ha-1, resulting in a total surface

soil C pool of 224 Tg C (Tg = g 9 1012; Table 4).

Stedman and Dahl (2008) reported a loss of 90,423 ha

of Wetlands in the northern Gulf of Mexico coastal

region between 1998 and 2004; this would translate to

a potential loss of 3.8 Tg C from these coastal wetland

soils (Table 4).

Discussion

Coastal wetlands contribute to the global carbon cycle

by sequestering carbon in soils and plant biomass and

by releasing greenhouse gases to the atmosphere. The

capacity of coastal wetlands to provide net carbon

reduction, therefore, requires that the rate of carbon

sequestration exceed the rate of carbon released to the

atmosphere (Whiting and Chanton 2001). While

coastal wetlands may be more valuable than other

ecosystems as carbon sinks (Choi and Wang 2004), in

the Gulf of Mexico, coastal wetlands are being lost at

an alarming rate (*25,000 ha year-1; Stedman and

Dahl 2008), consequently reducing their contribution

to the larger carbon mass balance of carbon seques-

tration and atmospheric carbon releases (DeLaune and

White 2012). Unfortunately very little is known about

the effects of wetland management on overall wetland

carbon dynamics (DeLaune and White 2012); how-

ever, management actions that reduce or reverse the

loss of coastal wetlands allow for those wetlands to

contribute to the regulation of the larger carbon budget

(Engle 2011; Li et al. 2004). Quantitative estimates of

carbon sequestration rates, carbon storage and green-

house gas emissions as well as models that link these

to wetland management and climate change mitigation

scenarios are needed to improve resource management

decisions (DeLaune and White 2012; Dumanski 2004;

Li et al. 2004; Powlson et al. 2011).

While this landscape-scale study contributes quan-

titative estimates of soil carbon pools and accumula-

tion rates in wetlands in the northern Gulf of Mexico

coastal region, several limitations must be addressed.Ta

ble

4E

stim

ated

soil

carb

on

sto

rag

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ls,

and

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mu

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ico

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tal

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lan

ds

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06

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09

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Tg

=1

012

g).

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ilca

rbo

np

oo

lsw

ere

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ula

ted

asp

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uct

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soil

carb

on

sto

rag

ean

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nd

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.A

nn

ual

soil

carb

on

accu

mu

lati

on

esti

mat

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ere

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ted

asth

ep

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uct

of

the

20

04

wet

lan

dar

eaes

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ates

and

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aver

age

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(mea

nan

dra

ng

esh

ow

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Wet

lan

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ass

Mea

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sity

(gC

cm-

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So

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(Mg

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19

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(gC

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20

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123

Our study measured carbon in the top 10–15 cm of soil

and, therefore, only provides estimates of surface

carbon density and accumulation. Surface soil carbon

accumulation, however, may not necessarily reflect

long-term burial and storage in coastal wetland soils

(Saintilan et al. 2013). Estimates of long-term carbon

storage in wetland soils would require that the top

meter of soil is used in order to incorporate the loss of

refractive carbon with depth. Vertical accretion rates

that are determined using different methods represent

different time scales (decades for 137Cs vs years for

feldspar marker) and should not be compared directly.

By choosing accretion rates that were determined

using the 137Cs method (except for palustrine forested

wetlands where only accretion rates using the feldspar

method were available), we tried to minimize vari-

ability due to different methods. However, even using

the same method, vertical accretion rates are variable

within wetland types, as shown for Louisiana salt

marshes in Table 5. Because we did not measure

vertical accretion directly as part of this study, we used

average accretion rates from the literature to calculate

potential soil carbon accumulation rates and applied

them as constants within each wetland sub-class. The

actual variability in soil carbon accumulation rates

within and across wetland types is likely much larger

than presented here.

Chmura et al. (2003) estimated soil OC densities

and accumulation rates in Gulf of Mexico tidal saline

marshes and mangrove wetlands from literature values

for bulk density, organic matter content, and accretion

rates. For most of the studies from the northern Gulf of

Mexico, OC content was estimated from organic

matter content using equations from Allen (1974) for

mangroves and from Craft et al. (1991) for salt

marshes [Chmura et al. 2003; note that the equation

from Craft et al. (1991) was reported incorrectly in

Chmura et al. (2003)—the correct equation is

OC = (0.4 9 LOI) ? (0.0025 9 LOI2)]. The equa-

tions used by Chmura et al. (2003) to convert organic

matter to organic carbon should only be applied to

estuarine marshes (Craft et al. 1991) and mangroves

(Allen 1974) as the equations were developed specif-

ically for these wetland types and there is no evidence

to suggest that they would be appropriate conversion

factors for other types of wetlands. We applied these

equations, therefore, only to the data from natural salt

and brackish marshes and mangroves in other studies

to estimate soil OC density for comparisons with our

study. In addition to the studies cited by Chmura et al.

(2003), we found several other relevant studies on soil

organic matter accumulation rates for estuarine and

palustrine wetlands in the northern Gulf of Mexico;

however, all but two of these studies (Craft and

Richardson 1993; Choi and Wang 2004) also only

reported organic matter content. While, our study

provides actual measured concentrations of both soil C

and OC, we could only compare soil OC density to

other studies in the literature. Literature values for soil

OC density were sparse for freshwater marshes and

forested wetlands; our study contributes new infor-

mation on the carbon content and potential accumu-

lation rates in these wetlands.

Fresh marshes in the northern Gulf of Mexico that

were not dominated by Cladium had soil OC densities

that ranged from 0.006–0.095 g OC cm-3. While no

comparative studies were available from the northern

Gulf of Mexico, Loomis and Craft (2010) reported

bulk density and OC content in tidal freshwater

wetland soils in Georgia. These wetlands were

Table 5 Comparison of soil OC density and accumulation rates from salt marshes in Louisiana. OC accumulation rates from studies

other than ours were calculated from organic matter (OM) content using the formula, OC = (0.4 9 OM) ? (0.0025 9 OM2) from

Craft et al. (1991) and reported accretion rates

Vertical accretion

rate (cm year-1)

Soil OC density

(g OC cm-3)

Soil OC accumulation

rate (g OC m-2 year-1)

Source

0.94 0.015–0.078 145–735 This study

0.35–1.13 0.010–0.019 41–186 Cahoon and Turner (1989)

1.10 0.028 309 Cahoon (1994)

0.59–1.40 0.023–0.028 139–378 Hatton et al. (1983)

0.47–0.68 0.021–0.024 96–161 DeLaune et al. (1989)

0.55–1.78 0.019–0.032 110–562 Nyman et al. (1993)

0.59–0.98 0.025–0.032 189–245 Nyman et al. (2006)


123

dominated by Zizaniopsis miliacea and bulk density

(0.23–0.27 g cm-3) and OC (9–12 %) were lower

than those from our study (Table 2); however, average

OC density (calculated from bulk density and %OC)

was equivalent to the average OC density for fresh

marshes from our study (0.03 g OC cm-3; Table 2).

In fresh marshes dominated by Cladium, soil OC

density (0.016–0.028 g OC cm-3) was lower than soil

OC density (0.037–0.045 g OC cm-3) reported by

Craft and Richardson (1993) for unenriched marshes

in the Everglades. The sites studied by Craft and

Richardson (1993), however, were located in the water

conservation area in the north–central Everglades

while our sites were located in Everglades National

Park.

Soil OC density in brackish marshes in the northern

Gulf of Mexico ranged from 0.01–0.08 g OC cm-3

which was slightly larger than the narrow range of OC

densities (0.016–0.031 g OC cm-3) calculated from

bulk density and organic matter content reported by

Cahoon (1994), DeLaune et al. (1989), Hatton et al.

(1983), and Nyman et al. (1993, 2006). The brackish

marshes in our study were located throughout the coast

of Louisiana while the literature studies were located

primarily in the Barataria–Terrebonne watersheds in

southeastern Louisiana (Cahoon 1994; Hatton et al.

1983; Nyman et al. 1993, 2006) and along Lake

Calcasieu in western Louisiana (DeLaune et al. 1989).

Our study reports higher soil OC density

([0.03 g OC cm-3) in brackish marshes in the At-

chafalaya–Vermilion–Teche watersheds and along the

Mississippi river floodplain.

Soil OC density in Gulf of Mexico salt marshes from

our study ranged from 0.009–0.078 g OC cm-3 which is

similar to the range of OC densities calculated from bulk

density and organic matter content reported by Hatton

et al. (1983; 0.021–0.028), Cahoon and Turner (1989;

0.010–0.019 g OC cm-3), DeLaune et al. (1989; 0.021–

0.024 g OC cm-3), Nyman et al. (1993; 0.019–0.032

g OC cm-3), Alford et al. (1997; 0.037 g OC cm-3),

Callaway et al. (1997; 0.027–0.040 g OC cm-3) and

Nyman et al. (2006; 0.025–0.032 g OC cm-3) and

measured OC densities (0.025–0.052 g OC cm-3)

reported by Choi and Wang (2004).

Soil OC density in mangrove soils in Florida ranged

from 0.016 to 0.040 g OC cm-3. Callaway et al. (1997)

and Cahoon and Lynch (1997) reported bulk density and

organic matter content from mangrove soils in the

Florida Keys and Rookery Bay, Florida; calculated OC

densities (0.04–0.07 g OC cm-3) from these studies

were slightly higher than OC densities from mangrove

soils in our study. Mangrove sites in our study, however,

were located throughout southwest Florida, with only

two sites in the Florida Keys and one site near Rookery

Bay, Florida.

Soil OC accumulation rates calculated from OC

densities from our study and vertical accretion rates

from the literature ranged from 46 to 627 g OC m-2 -

year-1 for brackish marshes and from 49 to

735 g OC m-2 year-1 for salt marshes. These rates

are within the range of soil OC accumulation rates for

Gulf of Mexico salt marshes that were estimated by

Chmura et al. (2003). DeLaune and White (2012)

reported soil OC accumulation rates that were esti-

mated from studies that reported soil organic matter

accumulation rates for Louisiana marshes (DeLaune

and Pezeshki 2003; Hatton et al. 1983; Nyman

et al. 2006). These estimates ranged from 132 to

338 g OC m-2 year-1 for brackish marshes and from

237 to 346 g OC m-2 year-1 for salt marshes (DeL-

aune and White 2012; DeLaune and Pezeshki 2003;

Hatton et al. 1983; Nyman et al. 2006). Smith et al.

(1983a) reported a range of 183–296 g OC m-2 -

year-1 for salt and brackish marshes in Louisiana. All

but two of the brackish marsh sites in our study were

located in Louisiana, yet, while our study shows an

overall greater range of soil OC accumulation rates for

brackish marshes than previous studies, if the highest

and lowest estimates are omitted, the range of soil OC

accumulation rates from brackish marshes in our study

would be 128–539 g OC m-2 year-1. Soil OC accu-

mulation rates for salt marshes in Louisiana from our

study also showed a greater range (145–735 g

OC m-2 year-1) than the average estimates presented

by DeLaune and White (2012). Part of the reason for

this difference could be that DeLaune and White

(2012) calculated OC accumulation rates from organic

matter density (multiplying by 0.56) and presented

only the average rates. In addition, the vertical

accretion rates in the studies cited by DeLaune and

White (2012) ranged from 0.59 to 1.35 cm year-1,

whereas we applied a constant accretion rate that

represented the average accretion rate from multiple

studies. Our soil OC accumulation rates for Louisiana

salt marshes are more similar to the ranges of soil OC

accumulation rates calculated from other studies

(41–562 g OC m-2 year-1), using the conversion

factor from Craft et al. (1991) (Table 5), Choi and


123

Wang (2004) reported soil C accumulation rates in

Florida salt marshes dominated by J. roemerianus,

ranging from 18 to 193 g C m-2 year-1. Three salt

marsh sites from our study were also located in Florida

and dominated by J. roemerianus. Soil OC accumu-

lation at these sites ranged from 173–446 g OC m-2 -

year-1, which was higher than those reported by Choi

and Wang (2004).

Our estimated soil OC accumulation rates showed

low variability across all mangrove sites from Rook-

ery Bay, Florida Keys, and Everglades National Park,

and the range for mangroves (42–104 g OC m-2 -

year-1) was lower than values reported in the litera-

ture. Chmura et al. (2003) calculated soil OC

accumulation rates in Florida mangrove wetlands

from carbon density and accretion rates reported by

Cahoon and Lynch (1997) and Callaway et al. (1997).

The estimated soil OC accumulation rates using data

from Cahoon and Lynch (1997) ranged from 222 to

381 g OC m-2 year-1 and were based on accretion

rates (0.44–0.78 cm year-1) determined using the

feldspar marker method. We used an average accretion

rate of 0.26 cm year-1 (Table 1), determined using137Cs method (Callaway et al. 1997; Lynch et al.

1989), which resulted in lower estimates of soil OC

accumulation rates for the mangrove sites in our study.

Craft and Richardson (1993) reported soil OC

accumulation rates ranging from 54 to 130 g OC

m-2 year-1 for unenriched, Cladium-dominated,

fresh marshes in the central Everglades. Compared

to our two sites in Everglades National Park, Craft and

Richardson’s (1993) sites had lower bulk density

(0.08–0.11 vs 0.27–0.38 g cm-3 at our sites) and

higher TOC content (39–50 vs 5–6 % TOC at our

sites).

OC accumulation rates in fresh marsh soils in

Louisiana ranged from 219 to 301 g OC m-2 year-1

(DeLaune and White 2012; Hatton et al. 1983; Nyman

et al. 2006; Smith et al. 1983a). Soil OC accumulation

rates for non-Cladium fresh marshes from our study

ranged from 48 to 796 g OC m-2 year-1, which is a

greater range than reported in the literature. Several

studies conducted across salinity gradients in Louisi-

ana marshes showed no appreciable differences in soil

OC accumulation between fresh and salt marshes

(Hatton et al. 1983; Nyman et al. 2006; Smith et al.

1983a). Our results showed a similar pattern with no

significant differences in the average soil OC accu-

mulation rates for non-Cladium fresh marshes,

brackish, and salt marshes (Table 3). In contrast, tidal

freshwater and brackish marshes in Georgia were

reported to have significantly higher OC accumulation

rates than salt marshes (Loomis and Craft 2010).

Cypress–tupelo forested wetlands had the highest

average rate of soil OC accumulation

(372 g OC m-2 year-1) based on our estimates for

wetlands in the northern Gulf of Mexico coastal region

(Table 3). Our estimate for soil OC accumulation in

bottomland hardwood forested wetlands was much

lower (141 g C m-2 year-1). We could not find any

comparable estimates for coastal forested wetlands in

the literature; coastal forested wetlands are usually

included in estimates for forests, in general. While our

estimates for soil carbon accumulation in coastal

forested wetlands are high, soil carbon sequestration in

coastal forested wetlands may be limited and tempo-

ral, depending on wetland type and hydrological

regime (Yu et al. 2006). For forested wetlands to be

carbon sinks over the long-term, stable hydrologic

conditions are necessary to sustain carbon sequestra-

tion (Trettin and Jurgensen 2003).

Quantifying carbon sequestration rates in coastal

wetland soils is only a first step to understanding the

contribution of coastal wetlands to ameliorate green-

house gas emissions. Assessment of the potential

value of coastal wetlands as carbon sinks in the future

must also include consideration of greenhouse gas

emissions (Chmura et al. 2003; Mitsch et al. 2012;

Whiting and Chanton 2001). Determining whether

coastal wetlands ultimately function as net carbon

sources or sinks requires balancing carbon storage in

soils and plant biomass with releases of CH4 and N2O.

While salt marshes and mangroves have negligible

CH4 emissions because of sulfate inhibition, signifi-

cant, but highly variable, CH4 emissions have been

measured in fresh and brackish coastal wetlands in the

Mississippi River delta (Alford et al. 1997) and in the

Florida Everglades (Bartlett et al. 1989; Harriss et al.

1988). DeLaune et al. (1990) estimated CH4 emissions

from coastal wetlands in the northern Gulf of Mexico

to be 1.5 9 1012 g CH4–C year-1, while Smith et al.

(1983b) estimated that coastal wetlands in the northern

Gulf of Mexico contribute 3.3 9 109 g N2O. Mitsch

et al. (2012) suggest that while many wetlands release

significant amounts of CH4, most temperate and

tropical wetlands are still net carbon sinks.

In this paper, we expand current knowledge about

the potential for wetlands in the northern Gulf of


123

Mexico coastal region to sequester carbon by report-

ing measured soil carbon densities and estimated soil

carbon accumulation rates for a wide range of wetland

types. We recognize that these estimates have large

uncertainties and do not adequately capture the

inherent spatial and temporal variability within and

among wetland types. This unknown variability within

wetland types can introduce significant uncertainty in

extrapolations to the regional scale (e.g., Bartlett et al.

1989). We believe, however, that these estimates will

improve our ability to incorporate the value of

wetlands in the northern Gulf of Mexico coastal

region in future wetland management decisions

related to climate change. As climate change policies

are implemented with the development of carbon sinks

to offset greenhouse gas emissions, the protection,

restoration, and management of coastal wetlands to

sequester carbon may be considered for carbon credits

in the future (Whiting and Chanton 2001). The carbon

storage potential in coastal wetlands should be

included in carbon accounting and inventories, devel-

opment of financial incentive mechanisms, and

amendment of policies to reduce loss of these ecosys-

tems. The use of mitigation wetlands could be

carefully designed and managed to influence carbon

uptake and minimize release of greenhouse gases, and

the value of carbon sequestration in coastal wetlands

could be factored into a cost/benefit analysis of coastal

restoration options (DeLaune and White 2012; Whit-

ing and Chanton 2001). The total value of carbon

sequestered by Louisiana coastal wetlands, for exam-

ple, has been estimated to be $29.7–$44.5 mil-

lion year-1 (DeLaune and White 2012). To provide

the best information to policy makers on the contri-

bution of carbon sequestration in coastal wetlands to

climate change mitigation strategies, additional

research is needed to: (1) quantify carbon sequestra-

tion and greenhouse gas emissions; (2) develop carbon

flux and carbon accounting methodologies; and (3)

determine how different restoration and management

approaches influence carbon flux in coastal and near-

shore marine ecosystems (Powlson et al. 2011).

Acknowledgments We thank Alex Almario (USEPA), Tom

Heitmuller (USGS-retired), Darrin Dantin (US EPA), Pat

O’Donnell (Rookery Bay National Estuarine Research

Reserve), and the staff from the Louisiana Department of

Natural Resources Coastal Restoration Division, USGS-NWRC

Coastal Restoration Field Station (Baton Rouge) for conducting

the field work to assess the condition of wetlands in the northern

Gulf of Mexico coastal region in 2007 and 2008. Sincere

appreciation is also given to Amanda Nahlik, Matthew Harwell,

and the anonymous peer reviewers for their insightful comments

and recommendations to improve this manuscript. The

information in this document has been funded by the U.S.

Environmental Protection Agency. It has been subjected to

review by the National Health and Environmental Effects

Research Laboratory and approved for publication. The views

expressed in this paper are those of the authors and do not

necessarily reflect the views or policies of the U.S. Environmental

Protection Agency. Mention of trade names or commercial

products does not constitute endorsement or recommendation for

use. This is contribution number 1444 from the U.S. EPA, Office

of Research and Development, National Health and

Environmental Effects Laboratory, Gulf Ecology Division.

References

Alford DP, DeLaune RD, Lindau CW (1997) Methane flux from

Mississippi river deltaic plain wetlands. Biogeochemistry

37:227–236

Allen SE (1974) Chemical analysis of ecological materials.

Blackwell Scientific, London

Bartlett DS, Bartlett KB, Harman JM et al (1989) Methane

emissions from the Florida Everglades: patterns of vari-

ability in a regional wetland ecosystem. Glob Biogeochem

Cycles 3:363–374

Blake GR, Hartge KH (1986) Bulk density. In: Klute A (ed)

Methods of soil analysis. Part 1. Physical and mineralogi-

cal methods. American Society of Agronomy, Madison,

pp 363–375

Brantley CG, Day JW Jr, Lane RR et al (2008) Primary pro-

duction, nutrient dynamics, and accretion of a coastal

freshwater forested wetland assimilation system in Loui-

siana. Ecol Eng 34:7–22

Bridgham SD, Megonigal JP, Keller JK, Bliss NB, Trettin C

(2006) The carbon balance of North American wetlands.

Wetlands 26:889–916

Bryant JC, Chabreck RH (1998) Effects of impoundment on

vertical accretion of coastal marsh. Estuaries 21:416–422

Cahoon DR (1994) Recent accretion in two managed marsh

impoundments in coastal Louisiana. Ecol Appl 4:166–176

Cahoon DR, Lynch JC (1997) Vertical accretion and shallow

subsidence in a mangrove forest of southwestern Florida,

USA. Mangroves Salt Marshes 1:173–186

Cahoon DR, Turner RE (1989) Accretion and canal impacts in a

rapidly subsiding wetland II. Feldspar marker horizon

technique. Estuaries 12:260–268

Callaway JC, DeLaune RD, Patrick WH Jr (1997) Sediment

accretion rates from four coastal wetlands along the Gulf of

Mexico. J Coast Res 13:181–191

Chmura GL, Anisfeld SC, Cahoon, Lynch JC (2003) Global

carbon sequestration in tidal, saline wetland soils. Glob

Biogeochem Cycles 17:22.1–22.12

Choi Y, Wang Y (2004) Dynamics of carbon sequestration in a

coastal wetland using radiocarbon measurements. Glob Bio-

geochem Cycles 18:GB4016. doi:10.1029/2004GB002261

Cowardin LM, Carter V, Golet FC, LaRoe ET (1979) Classifi-

cation of wetlands and deepwater habitats of the United

States. US Fish and Wildlife Service, Washington


123

http://dx.doi.org/10.1029/2004GB002261

Craft CB, Richardson CJ (1993) Peat accretion and N, P, and

organic C accumulation in nutrient-enriched and unen-

riched Everglades peatlands. Ecol Appl 3:446–458

Craft CB, Seneca ED, Broome SW (1991) Loss on ignition and

Kjeldahl digestion for estimating organic carbon and total

nitrogen in estuarine marsh soils: calibration with dry

combustion. Estuaries 14:175–179

Craft CB, Seneca ED, Broome SW (1993) Vertical accretion in

microtidal regularly and irregularly flooded estuarine

marshes. Estuar Coast Shelf Sci 37:371–386

Dahl TE, Bergeson MT (2009) Technical procedures for con-

ducting status and trends of the Nation’s wetlands. Division

of Habitat and Resource Conservation, US Fish and

Wildlife Service, Washington, p 74

Day JW, Shaffer GP, Britsch LD et al (2000) Pattern and process

of land loss in the Mississippi delta: a spatial and temporal

analysis of wetland habitat change. Estuaries 23:425–438

DeLaune RD, Pezeshki SR (2003) The role of soil organic

carbon in maintaining surface elevation in rapidly subsid-

ing US Gulf of Mexico coastal marshes. Water Air Soil

Pollut 3:167–179

DeLaune RD, White JR (2012) Will coastal wetlands continue

to sequester carbon in response to an increase in global sea

level?: a case study of the rapidly subsiding Mississippi

river deltaic plain. Clim Change 110:297–314

DeLaune RD, Patrick WH Jr, Buresh RJ (1978) Sedimentation

rates determined by 137Cs dating in a rapidly accreting salt

marsh. Nature 275:532–533

DeLaune RD, Baumann RH, Gosselink JG (1983) Relationship

among vertical accretion, coastal submergence and erosion in

a Louisiana Gulf coast marsh. J Sediment Petrol 53:146–157

DeLaune RD, Whitcomb JH, Patrick WH Jr, Pardue JH, Pez-

eshki SR (1989) Accretion and canal impacts in a rapidly

subsiding wetland. I. 137Cs and 210Pb techniques. Estuaries

12:247–259

DeLaune RD, Patrick WH Jr, Van Breemen N (1990) Processes

governing marsh formation in a rapidly subsiding coastal

environment. Catena 17:277–288

Dumanski J (2004) Carbon sequestration, soil conservation, and

the Kyoto Protocol: summary of implications. Clim

Change 65:255–261

Engle VD (2011) Estimating the provision of ecosystem ser-

vices by Gulf of Mexico coastal wetlands. Wetlands

31:179–193

Fennessy, Jacobs AD, Kentula ME (2004) Review of rapid

methods for assessing wetland condition. EPA/620/R-04/

009. US Environmental Protection Agency, Washington,

DC

Fennessy MS, Jacobs AD, Kentula ME (2007) An evaluation of

rapid assessment methods for assessing condition of wet-

lands. Wetlands 27:543–560

Field DW, Reyer AJ, Genovese PV, Shearer BD (1991) Coastal

wetlands of the United States: an accounting of a valuable

national resource. National Oceanic and Atmospheric

Administration, National Ocean Service, Maryland

Harriss RC, Sebacher DI, Bartlett KB, Bartlett DS, Crill PM

(1988) Sources of atmospheric methane in the south Flor-

ida environment. Glob Biogeochem Cycles 2:231–243

Hatton RS, DeLaune RD, Patrick WH Jr (1983) Sedimentation,

accretion, and subsidence in marshes of Barataria Basin,

Louisiana. Limnol Oceanogr 28:494–502

Jobbagy EG, Jackson RB (2000) The vertical distribution of soil

organic carbon and its relation to climate and vegetation.

Ecol Appl 10:423–436

Karam A (1993) Chemical properties of organic soils. In: Carter

MR (ed) Soil sampling and methods of analysis. Lewis

Publishers, Boca Raton

Li C, Cui J, Sun G, Trettin C (2004) Modeling impacts of

management on carbon sequestration and trace gas emis-

sions in forested wetland ecosystems. Environ Manag

33(Suppl 1):S176–S186

Loomis MJ, Craft CB (2010) Carbon sequestration and

nutrient (nitrogen, phosphorus) accumulation in river-

dominated tidal marshes, Georgia, USA. Wetl Soils

74:1028–1036

Lynch JC, Meriwether JR, McKee BA, Vera-Herrera F, Twilley

RR (1989) Recent accretion in mangrove ecosystems based

on 137Cs and 210Pb. Estuaries 12:264–299

Markewich HW, Britsch LD, Buell GR et al (1998) Carbon

storage and late holocene chronostratigraphy of a Missis-

sippi river deltaic marsh, St. Bernard Parish, Louisiana. US

Geological Survey Open-File Report, p 98–36

Mcleod E, Chmura GL, Bouillon S et al (2011) A blueprint for

blue carbon: toward an improved understanding of the role

of vegetated coastal habitats in sequestering CO2. Front

Ecol Environ 9:552–560

Mitsch WJ, Bernal B, Nahlik AM et al (2012) Wetlands, carbon,

and climate change. Landsc Ecol 28:583–597

Nestlerode JA, Engle VD, Bourgeois P, Heitmuller PT, Ma-

cauley J, Allen Y (2009) An integrated approach to assess

broad-scale condition of Gulf of Mexico coastal wetlands.

Environ Monit Assess 150:21–29

Nyman JA, DeLaune RD, Roberts HH, Patrick WH Jr (1993)

Relationship between vegetation and soil formation in a

rapidly submerging coastal marsh. Mar Ecol Prog Ser

96:269–279

Nyman JA, DeLaune RD, Pezeshki SR, Patrick WH Jr (1995)

Organic matter fluxes and marsh stability in a rapidly

submerging estuarine marsh. Estuaries 18:207–218

Nyman JA, Walters RJ, DeLaune RD, Patrick WH Jr (2006)

Marsh vertical accretion via vegetative growth. Estuar

Coast Shelf Sci 69:370–380

Poffenbarger HJ, Needelman BA, Megonigal JP (2011) Salinity

influence on methane emissions from tidal marshes. Wet-

lands 31:831–842

Post WM, Izaurralde RC, Jastrow JD et al (2004) Enhancement

of carbon sequestration in US soils. Bioscience

54:895–908

Powlson DS, Whitmore AP, Goulding KWR (2011) Soil carbon

sequestration to mitigate climate change: a critical re-

examination to identify the true and the false. Eur J Soil Sci

62:42–55

Rabenhorst MC (1995) Carbon storage in tidal marsh soils. In:

Lal R, Kimble J, Levine E, Stewart BA (eds) Soils and

global change. CRC Press, Florida, pp 93–104

Rybczyk JM, Day JW Jr, Conner WH (2002) The impact of

wastewater effluent on accretion and decomposition in a

subsiding forested wetland. Wetlands 22:18–32

Saintilan N, Rogers K, Mazumder D, Woodroffe C (2013) Al-

lochthonous and autochthonous contributions to carbon

accumulation and carbon store in southeastern Australian

coastal wetlands. Estuar Coast Shelf Sci 128:84–92


123

Smith CJ, DeLaune RD, Patrick WH Jr (1983a) Carbon dioxide

emission and carbon accumulation in coastal wetlands.

Estuar Coast Shelf Sci 17:12–29

Smith CJ, DeLaune RD, Patrick WH Jr (1983b) Nitrous oxide

emission from Gulf Coast wetlands. Geochim Cosmochim

Acta 47:1805–1814

Stedman S, Dahl TE (2008) Status and trends of wetlands in the

coastal watersheds of the eastern United States 1998–2004.

National Oceanic and Atmospheric Administration,

National Marine Fisheries Service and US Department of

the Interior, Fish and Wildlife Service, Maryland

Stevens DL Jr, Olsen AR (2004) Spatially balanced sampling of

natural resources. J Am Stat Assoc 99:262–278

Tiessen H, Moir JO (1993) Total and organic carbon. In: Carter

MR, for Canadian Society of Soil Science (eds). Soil

sampling and methods of analysis. Lewis Publishers,

London, pp 187–199

Trettin CC, Jurgensen MF (2003) Carbon cycling in wetland

forest soils. In: Kimble JM et al (eds) Potential of US forest

soils to sequester carbon and mitigate the greenhouse

effect. CRC Press, Florida, pp 311–346

Turner RE (1997) Wetland loss in the northern Gulf of Mexico:

multiple working hypotheses. Estuaries 20:1–13

USEPA (2006) Application of elements of a state water moni-

toring and assessment program for wetlands. Office of

Wetlands, Oceans and Watersheds, Wetlands Division, US

Environmental Protection Agency, Washington

USEPA (2011) Inventory of US greenhouse gas emissions and

sinks: 1990–2009. EPA 430-R-11-005. US Environmental

Protection Agency, Washington

Whiting GJ, Chanton JP (2001) Greenhouse carbon balance of

wetlands: methane emission versus carbon sequestration.

Tellus 53B:521–528

Yu K, Faulkner SP, Patrick WH Jr (2006) Redox potential

characterization and soil greenhouse gas concentration

across a hydrological gradient in a Gulf coast forest.

Chemosphere 62:905–914


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