NITROGEN ADVECTION AND DENITRIFICATION LOSS IN SOUTHEASTERN
NORTH CAROLINA SALT MARSHES
Matthew D. Lettrich
A Thesis Submitted to the
University of North Carolina Wilmington in Partial Fulfillment
of the Requirements for the Degree of
Master of Science
Center for Marine Science
University of North Carolina Wilmington
2011
Approved by
Advisory Committee
Lawrence B. Cahoon
Craig R. Tobias
Eric J. Henry
Chair
Accepted by
DN: cn=Robert D. Roer, o=UNCW,
ou=Dean of the Graduate School &
Research, email=roer@uncw.edu,
c=US
Date: 2011.07.18 16:30:02 -04'00'
______________________________
Dean, Graduate School
TABLE OF CONTENTS
ABSTRACT.....................................................................................................................................v
ACKNOWLEDGEMENTS ........................................................................................................... vi
LIST OF TABLES ........................................................................................................................ vii
LIST OF FIGURES ..................................................................................................................... viii
OVERVIEW ....................................................................................................................................1
CHAPTER 1. Advective Marsh-Estuary Exchange - Hydrologic and Chemical Fluxes ...............5
INTRODUCTION ...........................................................................................................................6
METHODS ......................................................................................................................................9
Site Description..............................................................................................................................10
Site Development ...........................................................................................................................10
Hydraulic Parameters .....................................................................................................................12
Porewater Transport .......................................................................................................................17
Chemical sampling.........................................................................................................................17
Marsh Scale Drainage - Whole Creek Tidal Flood vs. Ebb Flux Studies......................................18
RESULTS ......................................................................................................................................19
Tidal and Watershed Boundary Conditions ...................................................................................19
Porewater Drainage Drivers ...........................................................................................................27
Creekbank Drainage and Tides ..........................................................................................30
Creekbank Drainage and Upland Water Table ..................................................................30
Interior Drainage and Tides ...............................................................................................34
Interior Drainage and Upland Water Table .......................................................................44
Porewater Drainage Flux – Summary ............................................................................................44
ii
Porewater Nitrogen Flux from Marsh to Open Water ...................................................................47
Comparison of Darcy N Flux with Tidal N Exchange Estimates ..................................................49
DISCUSSION ................................................................................................................................55
Water and N Movement .................................................................................................................55
Water Movement-Creekbank .............................................................................................55
Water Movement-Interior ..................................................................................................56
N Export from Marsh to Open Water ................................................................................58
Controls on the Advective Fluxes ......................................................................................59
Predicting Drainage ...........................................................................................................61
Marsh Scale N fluxes - Darcy vs Tidal Approaches ......................................................................63
Spatial and Temporal Variance - Darcy.............................................................................63
Spatial and Temporal Variance – Tidal Flux .....................................................................64
Incomplete Capture of All Routes of N Exchange - Darcy ...............................................65
Incomplete Capture of All Routes of N Exchange – Tidal Flux........................................66
Scaling - Darcy ..................................................................................................................66
Scaling – Tidal Flux ...........................................................................................................67
CHAPTER 2. Marsh-Estuary Exchange – Attenuation of DIN Fluxes by Denitrification ..........68
Introduction
................................................................................................................................69
Methods..........................................................................................................................................71
Results ............................................................................................................................................75
Direct Denitrification .....................................................................................................................75
Coupled Denitrification – 15NO3 Addition ....................................................................................76
Coupled Denitrification – 15NH4 Addition ....................................................................................80
iii
Direct Denitrification Capacity ......................................................................................................82
Discussion ......................................................................................................................................85
Seasonal and Spatial Differences ...................................................................................................85
Comparisons to other systems .......................................................................................................86
Direct vs. Coupled Denitrification .................................................................................................88
Denitrification Capacity and Anammox ........................................................................................89
SUMMARY AND SYNTHESIS ...................................................................................................91
Scaling Magnitudes of Drainage and Denitrification ....................................................................92
Importance of Marsh Drainage and Denitrification Relative to Other Sources and Sinks. ...........93
Implications for Marsh N export in a Changing Climate ..............................................................94
LITERATURE CITED .................................................................................................................95
iv
ABSTRACT
Coastal wetlands serve as sources and sinks of nitrogen to surrounding estuarine
waters through advective drainage and denitrification. The advective nitrogen flux of three
intertidal estuary wetlands in the New River Estuary in North Carolina was determined using
Darcy-derived drainage measurements and calculating the difference between tidal ebb and
tidal flood flux. The magnitude of drainage was greatest and most closely linked to tidal
elevation in the most down-estuary site and was least in the up-estuary site ranging from a
daily mean drainage of 0.34 L m shoreline-1 day-1 in the up-estuary site to 87 L m shoreline-1
day-1 in the down-estuary site. Nitrogen concentrations in the marsh porewaters peaked in
late 2009. N flux was determined as a function of drainage (water flux) and porewater N
concentration. Advective N flux showed a seasonal pattern that increased in the summer and
the winter. Drainage was found to be correlated to tidal elevation within each site and
trended with tidal amplitude within the estuary, providing proxies for estimating advective N
flux at other sites when given those easily measured parameters combined with porewater N
concentration. Marsh denitrification was found to be generally greater in the creekbank than
in the interior and greater in May than in February. Coupled denitrification dominated direct
denitrification as ambient nitrate concentrations remained low.
Denitrification capacity
showed that with a nearly 1000-fold increase in nitrate concentration, NRE marshes could
maintain the rate of denitrification relative to ambient nitrate concentration.
v
ACKNOWLEDGEMENTS
This research was conducted under the Defense Coastal/Estuarine Research Program
(DCERP), funded by the Strategic Environmental Research and Development Program
(SERDP).
vi
LIST OF TABLES
Table
Page
1. Hydraulic conductivity (Kh) of Traps Bay, French Creek, and Freeman Creek. ..............26
2. Quarterly porewater concentrations of total dissolved nitrogen (TDN) based on
location within each marsh ................................................................................................48
3. Denitrification rates reported of other coastal systems ......................................................87
vii
LIST OF FIGURES
Figure
Page
1. Site map showing Marine Corps Base Camp Lejeune (a), French Creek (b), Traps
Bay (c), and Freeman Creek (d). Images are not to the same scale ..................................11
2.
Aerial photo of French Creek study site. White points represent piezometer clusters. ....13
3.
Aerial photo of Traps Bay study site. White dots represent piezometer clusters. ............14
4.
Aerial photo of Freeman Creek study site. White dots represent piezometer clusters. ....15
5.
Conceptual model of salt marsh nitrogen linkages. Tidal recharge supplies particulate
nitrogen (PN), dissolved inorganic nitrogen (DIN), and dissolved organic nitrogen
(DON). Groundwater (GW) inputs supply DIN. Burial removes PN. Denitrification
results in gaseous loss of N2. Porewaters drain to the adjacent estuary resulting in
export of DIN, DON, and PN from the salt marsh system. ...............................................16
6.
Daily maximum observed high tide (black) and low tide (grey) for French Creek (a),
Traps Bay (b), and Freeman Creek (c). Creekbank marsh surface elevation is
indicated by the dashed line...............................................................................................20
7.
Upland water table elevations for French Creek (a), Traps Bay (b), and Freeman
Creek (c). Freeman Creek and Traps Bay showed responses to precipitation events
while French Creek predominantly responded to tidal elevation. .....................................22
8.
Short duration fluctuations in upland water table elevation (grey), tidal elevation
(black), and precipitation (bar) for French Creek (a), Traps Bay (b), and Freeman
Creek (c). ...........................................................................................................................23
9.
Representative daily fluctuation of tidal elevation (grey), creekbank gradient
(triangle), and interior gradient (circle) for French Creek (a), Traps Bay(b), and
Freeman Creek (c). Dh represents the difference in hydraulic head between either the
tide gauge and creekbank piezometer (creekbank gradient), or the upland-marsh
border piezometer and the creekbank piezometer (interior gradient). Dx represents
the distance between the points where dh is determined. A positive gradient indicates
a slope from the marsh towards open water. .....................................................................25
10. Creekbank porewater elevation (triangles) and tidal elevation (black line) through a
daily tidal series for French Creek (a), Traps Bay (b), and Freeman Creek (c). Once
the marsh is flooded, tidal elevation is greater than porewater elevation, creating a
negative gradient. Shaded grey areas represent times of vertical recharge. .....................28
11. Vertical gradient for the creekbank (triangles) and marsh interior (circles) for French
Creek (a), Traps Bay (b), and Freeman Creek (c) during flooded periods of a
representative tidal cycle. Tidal elevation is indicated by the black line. Negative
gradients indicate vertical recharge of the marsh. .............................................................29
viii
12. Creekbank drainage versus tidal elevation for French Creek transect 3 (a), transect 5
(b), and transect 7 (c). A two-phase response was seen with increased drainage at
elevations below the marsh surface (black ) and recharge for tidal elevations greater
than the marsh surface (grey). Grey points greater than zero occur as the marsh is
draining on the falling tide and grey values less than zero occur as the marsh is
flooding on the rising tide. .................................................................................................31
13. Creekbank drainage versus tidal elevation for Traps Bay transect 1 (most upstream)
(a), transect 3 (b), transect 5 (c), and transect 7 (most downstream) (d). A two-phase
response was seen with increased drainage at elevations below the marsh surface
(black ) and recharge for tidal elevations greater than the marsh surface (grey). Grey
points greater than zero occur as the marsh is draining on the falling tide and grey
values less than zero occur as the marsh is flooding on the rising tide. ............................32
14. Creekbank drainage versus tidal elevation for Freeman Creek transect 1 (a), transect 3
(b), and transect 5 (c). A two-phase response was seen with increased drainage at
elevations below the marsh surface (black ) and recharge for tidal elevations greater
than the marsh surface (grey). Grey points greater than zero occur as the marsh is
draining on the falling tide and grey values less than zero occur as the marsh is
flooding on the rising tide. .................................................................................................33
15. Daily maximum creekbank drainage response to upland water table slope for French
Creek transect 3 (a), transect 5 (b), and transect 7 (c). ......................................................35
16. Daily maximum creekbank drainage response to upland water table slope for Traps
Bay transect 1 (a), transect 3 (b), transect 5 (c), and transect 7 (d).36
17. Daily maximum creekbank drainage response to upland water table slope for
Freeman Creek transect 1 (a), transect 3 (b), and transect 5(c). ........................................37
18. Interior marsh drainage response to tidal elevation for French Creek transect 3 (a),
transect 5(b), and transect 7 (c). A two-phase response was seen for tidal elevations
below interior marsh surface (black ) and above interior marsh surface (grey). ...............38
19. Interior marsh drainage response to tidal elevation for Traps Bay transect 1 (a),
transect 3 (b), transect 5 (c), and transect 7 (d). A three-phase response was seen for
tidal elevations below creekbank marsh surface (black ), between creekbank marsh
surface and interior marsh surface (light grey), and above interior marsh surface (dark
grey). ..................................................................................................................................39
20. Interior marsh drainage response to tidal elevation for Freeman Creek transect 1 (a),
transect 3 (b), and transect 5 (c). A three-phase response was seen for tidal elevations
below creekbank marsh surface (black ), between creekbank marsh surface and
interior marsh surface (light grey), and above interior marsh surface (dark grey). ...........40
21. Daily maximum interior drainage response to upland water table slope for French
Creek transect 3 (a), transect 5 (b), and transect 7(c). .......................................................41
ix
22. Daily maximum interior drainage response to upland water table slope for Traps Bay
transect 1 (a), transect 3 (b), transect 5 (c). and transect 7(d)............................................42
23. Daily maximum interior drainage response to upland water table slope for Freeman
Creek transect 1 (a), transect 3 (b), and transect 5(c). .......................................................43
24. All sites daily maximum drainage versus daily tidal amplitude. French Creek (open
circles), Traps Bay (black circles), and Freeman Creek (triangles). .................................45
25. Mean monthly drainage corrected for cross sectional drainage area for French Creek
(solid triangles), Traps Bay (open boxes), and Freeman Creek (solid diamonds). .........46
26. Monthly average nitrogen drainage for French Creek, Traps Bay, and Freeman Creek.
Freeman Creek drains two orders of magnitude more N per day than Traps Bay and
French Creek. .....................................................................................................................51
27. Monthly average Darcy-derived nitrogen flux for French Creek, Traps Bay, and
Freeman Creek. Freeman Creek drains two orders of magnitude more N per day than
Traps Bay and French Creek. Rates were scale up to the whole marsh using shoreline
length estimates of 1300m for French Creek, 900m for Traps Bay, and 9600m for
Freeman Creek ...................................................................................................................52
28. Traps Bay discharge L s-1 (gray area) (a,c), TDN concentration (black line) (a,c), and
tidal signature (b,d). Discharge out of the marsh is negative, discharge into the marsh
is positive. Tidal flux studies from September 10, 2009 (a,b) and May 3, 2010 (c,d)
are shown. ..........................................................................................................................53
29. Freeman Creek discharge L s-1 (gray area) (a,c), TDN concentration (black line)
(a,c), and tidal signature (b,d). Discharge out of the marsh is negative, discharge into
the marsh is positive. Tidal flux studies from September 11, 2009 (a,b) and May 4,
2010 (c,d) are shown..........................................................................................................54
30. Mean direct denitrification rates from 4 hour incubations for French Creek (a, b),
Traps Bay (c,d), and Freeman Creek (e, f). Black represents creekbank, grey
represents interior. .............................................................................................................77
31. Comparison of direct denitrification rates of May versus February (a) and marsh
interior versus creekbank (b). Triangles represent French Creek, boxes represent
Traps Bay, diamonds represent Freeman Creek. For May versus February (a), solid
symbols represent creekbank, open symbols represent interior. For interior versus
creekbank (b), solid symbols represent February, open symbols represent May. .............78
32. Mean coupled denitrification rates from 4 hour incubations for French Creek (a-d),
Traps Bay (e-h), and Freeman Creek (i-l). Black represents creekbank sediment
incubations, grey represents interior sediment incubations. ..................................................
33. Comparison of coupled denitrification rates of marsh May versus February (a,b) and
interior versus creekbank (c,d). Triangles represent French Creek, boxes represent
x
Traps Bay, diamonds represent Freeman Creek. For May versus February (a,b), solid
symbols represent creekbank, open symbols represent interior. For interior versus
creekbank (c,d), solid symbols represent February, open symbols represent May. ..........81
34. Direct denitrification versus coupled denitrification for all samplings. Triangles
represent French Creek, boxes represent Traps Bay, diamonds represent Freeman
Creek. Closed symbols represent February samples, open symbol represent May
samples. .............................................................................................................................83
35. Denitrification and ANAMMOX capacity for French Creek creekbank (a) and interior
(b), Traps Bay creekbank (c) and interior (d), and Freeman Creek creekbank (e) and
interior (f). Solid symbols denote denitrification, open symbols denote
ANAMMOX.undetectable at ambient concentrations and ranged from 0-<2% of the
denitrification at ambient concentrations and at higher NO3 treatments, respectively. ....84
xi
OVERVIEW
Rates of nutrient loading and turnover impact water quality, fisheries, and recreational
uses of coastal waters (Valiela et al. 1990, Cloern 2001). Dissolved inorganic nitrogen (DIN) is
the principal nutrient limiting estuarine primary production, but in excess leads to eutrophication,
hypoxia, algal blooms, and altered ecosystem structure (Howarth et al. 2000). The availability of
DIN is controlled by external rates of loading from the watershed and rates of DIN
recycling/removal within and between ecosystem components within the estuarine landscape.
Specific habitats that can modify both incoming loads of DIN from the watershed and also affect
turnover of DIN once it enters the estuary play a central role in regulating overall DIN
availability on an estuarine scale. Intertidal marshes represent such a habitat. They are situated
between the nitrogen rich upland and the nitrogen-limited estuary; well-positioned to impact the
DIN budget through marsh-upland interactions and through marsh-estuary interactions.
Much effort has been exerted over the past 35 years to generally characterize the source
and sink nature of intertidal marshes with respect to nitrogen (Valiela and Teal 1979, Correll
1981, Childers et al. 1994). Past studies have relied on whole-creek tidal exchange or flume
approaches, which are limited to certain marsh geomorphologies (Spurrier and Kjerfve 1988,
Dame 1994, Childers et al. 2000). No clear unifying consensus has emerged over whether
marshes uniformly behave in all locations at all times as a source or sink. There is some
evidence that marsh import or export of N depends on N speciation, tidal range, and marsh age.
There remains considerable inter- and intra-marsh spatial variability in the direction and
magnitudes of marsh N fluxes. It is possible that alternate approaches that operate on scales
smaller than whole tidal creeks could prove valuable for addressing the source or sink dynamics
on the scales at which they vary and extend the work to other marsh geomorphologies not
amenable to whole-system approaches.
Groundwater, tidal exchange, burial, and denitrification represent significant pathways
through which marshes may regulate N speciation and availability in adjacent estuarine waters
(Tobias and Neubauer 2009). Specifically for DIN, groundwater discharge and tidal delivery
during flooding serve as principal sources of N to the marsh, while denitrification and porewater
drainage serve as the primary modes of N export from the marsh (Howes et al. 1996). DIN
exchanges regulated by groundwater, tidal infiltration, and drainage are integrally linked to
fluxes of water (i.e. advection). Groundwater discharge and tidal infiltration both contribute to
the total volume of porewater available for drainage. Quantifying the total “water” flux of
porewater through the rhizoshpere is the requisite first step for estimating the chemical fluxes of
DIN and dissolved organic nitrogen (DON).
The drainage flux of porewater integrates
contributions from groundwater, tidal infiltration, and precipitation, with water loss from
macrophyte evapotransipiration.
During drainage through the rhizosphere, porewater may
acquire additional DIN/DON owed to high rates of organic matter mineralization, which leads to
the typically N-rich character of marsh porewaters. Coupling the porewater drainage “water”
flux with measurements of the N chemical composition of the porewater permits calculation of
the DIN/DON chemical flux from the marsh to adjacent open water.
The magnitude of both
the drainage water and N fluxes varies temporally from hours to months as the magnitude of the
components of the water budget changes in response to tidal cycles, precipitation patterns, and
temperature, and mineralization rates change with growing season.
Second only to accretion of marsh sediment, denitrification represents one of the largest
marsh sinks for total nitrogen and is typically on par with rates of plant uptake for DIN removal.
2
Unlike plant uptake or accretion that stores N in the marsh on varying timescales, denitrification
converts N into N2 gas and facilitates its complete removal from the coastal landscape to the
atmosphere. Unlike the drainage N flux, denitrification is uncoupled from advective water
transport. While denitrification can be affected by patterns of tidal infiltration and drainage,
rates of denitrification can be calculated independently of water movement. Marsh denitrification
operates either directly on nitrate supplied by flooding water, precipitation, and groundwater
(Seitzinger et al. 2006) or “indirectly” on nitrate produced from the nitrification of ammonium
produced during mineralization of organic matter in marsh sediments (“coupled denitrification”;
Hammersley and Howes 2005, Seitzinger et al. 2006). It therefore has the potential to attenuate
N availability by acting on allochthonous N loads and by decreasing the amount of N internally
recycled within the marsh that would otherwise be available for drainage.
The net effect of the groundwater nitrogen inputs, tidal infiltration and drainage, and
denitrification ultimately modify the availability of N to the adjacent estuary and resultant water
quality. The overall balance of these processes determines the role and magnitude of intertidal
marshes within the landscape as a source or sink for DIN.
The body of research contained herein examines two fundamental pathways and
processes through which intertidal marshes modify nitrogen cycling in the coastal marine
landscape. First, chapter 1 considers the combined interactions of the marsh, watershed, and
estuary in the context of advective fluxes of water from the marsh to the estuary. Focus is given
to the role of porewater drainage as a primary conduit for DIN/DON transport to estuarine
waters. Second, chapter 2 considers N removal in the marsh via denitrification; a process
independent of water exchanges yet a typically important DIN removal process in marshes.
Determining the balance of these source and sink terms is necessary for assessing the importance
3
of intertidal marshes on response and resilience of coastal ecosystems to exogenous nitrogen
inputs.
4
CHAPTER 1. ADVECTIVE MARSH-ESTUARY EXCHANGE - HYDROLOGIC AND
CHEMICAL FLUXES
INTRODUCTION
Intertidal marshes are characterized by alternating tidal wetting and draining. Marsh
porewaters are derived from a mixture of groundwater, tidal infiltration (recharge), and
precipitation, and are modified by water losses due to evapotranspiration. Porewater chemistry
is controlled by contributions from these various sources of water and by in-situ N cycling in the
subsurface, most notably organic matter mineralization which results in typically high
concentrations of both DIN and DON (Valiela et al. 1978, Anderson et al. 1997). Porewaters can
drain to the adjacent surface waters through seepage during periods of low water or diffuse into
surface waters when the marsh is flooded (Harvey et al. 1995, Gardner 2005).
Groundwater input is focused at the marsh-upland interface, being driven by the water
table height in the watershed (Harvey and Odum 1990). Typically the shallow aquifer is the
primary source (Tobias et al. 2001b), however there are exceptions in which local geology
permits the deep aquifer to contribute. While previous studies show that seasonal effects vary by
location (Valiela et al. 1978, Tobias et al. 2001b), groundwater input is a function of water table
height regardless of location. Groundwater discharge to surrounding waters has been found to be
low when compared to freshwater runoff, accounting for ~10% of the total freshwater input
(Portnoy et al. 1997, Bowen et al. 2007). However, if the shallow aquifer is high in N, it can
supply the salt marsh with large amounts of allochthanous N.
With elevated nutrient
concentrations, groundwater discharge may play a large part in the landscape level loading of
nutrients to the adjacent estuary (Gardner 1975, Valiela et al. 1978, Corbett et al. 1999). In some
cases the shallow aquifer is N-poor, but groundwater input still serves as a major component in
the water budget that promotes drainage by raising the hydraulic gradient from the upland border
to the creekbank (Tobias et al. 2001b).
6
Tidal infiltration occurs at the marsh-estuary interface and is derived from a pattern of
tidal flooding and draining. Along the east coast of the United States, the tidal signature is
dominated by a semi-diurnal lunar tide and a semimonthly spring-neap signature that modifies
the daily tides. Tidal signatures are further modified stochastically by storm systems and Ekman
tides propagated by offshore wind patterns (Bacopoulos et al. 2009). Due to these variations, the
magnitude of infiltration deviates from a regular periodicity but does not deviate seasonally like
groundwater inputs. In addition to the frequency of flooding, magnitude is also a function of the
thickness of the unsaturated zone and the specific yield of the marsh rhizosphere (Harvey and
Odum 1990, Tobias et al 2001b). Infiltration of precipitation follows the same functions but
occurs only during periods of heavy rainfall and is typically a small component of the subsurface
water budget. The accepted model for infiltration is one of vertical recharge when the marsh is
flooded (Harvey et al. 1987). The model is based on the premise that hydraulic conductivity of
the saturated zone inhibits horizontal infiltration and that once the rising tide overtops the marsh
surface, vertical infiltration of the unsaturated zone rapidly occurs.
Groundwater discharge along the upland marsh edge mixes with infiltrating tidal water to
facilitate the drainage of nutrient-rich porewaters to surrounding surface water (Harvey and
Odum 1990, Howes and Goehringer 1994, Howes et al. 1996, Tobias et al. 2001b). The marsh
drains on two functional timescales. First, there is slow and constant drainage from the back of
the marsh to the creekbank represented by the slope of the water table within the marsh (Harvey
et al. 1987, Tobias et al. 2001b). Second, there is rapid drainage in the creekbank zone resulting
from draining and filling with each flooding event (Harvey et al. 1987, Howes and Goehringer
1994, Harvey et al. 1995). The two timescales combine to create a system that includes slow
transport of porewater from the back of the marsh to the creekbank zone which is then rapidly
7
drained and recharged. While input may be horizontal (groundwater) or vertical (groundwater,
tidal infiltration, precipitation), drainage is a horizontal export flux (Harvey et al. 1987, Howes
and Goehringer 1994, Tobias et al. 2001b).
Several approaches have been used to estimate marsh N export ranging from wholesystem approaches to small scale measurements. Radium balance approaches that estimate water
flux based on the half life of radium isotopes have been successfully implemented (Krest et al.
2000, Charette et al. 2003) for large marshes but are expensive, do not necessarily account for
other nutrient sources, and do not correlate directly to the water or nutrient budgets. Whole
creek studies that sample incoming and outgoing waters throughout a tidal cycle (Valiela et al.
1978, Woodwell et al. 1979, Roman 1984, Wolaver et al. 1988) produce a net flux for the marsh
but are reliant on the presence of a centralized tidal creek. The studies are labor-intensive and
can only be performed at specific times, providing fine temporal resolution but for only a very
brief time and limiting transferability to other systems and times.
Smaller scale approaches focus on transects or plots within the marsh. The salt balance
approach (Harvey and Odum 1990, Tobias et al. 2001b) works well in areas of fresh
groundwater and has less error in calculation than other approaches but is less effective when
used in brackish waters that are common in intertidal marshes. Flume studies (Childers 1994,
Childers et al. 2000) and flux chamber studies (Chambers et al. 1992, Windham-Myers 2005)
offer integrated measurements of a specific marsh surface and the ability to isolate individual
aspects of the marsh. The protocols are easily transferred between sites but require intensive
intstrumentation to characterize a single site.
Hydrogeologic techniques applied to marsh hydrology have also been used to provide
estimates for porewater drainage. Hydraulic head (Darcy) calculations use direct measurements
8
of the forcing functions but assume sediment homogeneity which may increase error (Tobias et
al. 2001b) and the calculations may not capture processing at the discharge interface or diffusion
at the marsh surface.
Porewater solute concentrations are highly variable between and within marshes due to
vegetation regimes (Windham-Myers 2005), microbial processes (Seitzinger et al. 2006), and
source waters (Harvey et al. 1995). In general, with the exception of very young marshes
(Osgood and Zieman 1993a, Osgood and Zieman 1993b), porewaters are richer in DIN and DON
than tidal flooding water. Consequently, marsh drainage constitutes a net flux of dissolved
reactive N to adjacent coastal waters.
The objective of this study was to assess the marsh to estuary flux of dissolved N. The
work relies primarily on small-scale, high-frequency hydraulic head/gradient measurements
coupled to porewater chemistry characterization. A long-term continuous record of porewater
drainage from three sites within an estuary provides temporal resolution on a monthly, daily, and
hourly scale. Fine temporal and spatial resolution was used to calibrate porewater drainage to
easily measured physical drivers at the marsh-estuary and marsh-upland interfaces. Small-scale
hydraulic gradient-based drainage N flux estimates are scaled up and compared to marsh-scale
tidal N flux of select marshes.
METHODS
Two approaches were used to quantify marsh export of DIN and DON. First, small-scale
hydraulic gradient induced fluxes were calculated using measurements of hydraulic gradient and
conductivity (the Darcy method). Second, whole-creek tidal flood and ebb flux studies were
used on a subset of marshes. In all, three marshes were studied in the New River Estuary, NC
that span a gradient in salinity and tidal range.
9
Site Description
Freeman’s Creek, Traps Bay, and French Creek are three intertidal marshes located on
Marine Corps Base Camp Lejeune in Jacksonville, NC along an estuarine salinity gradient in the
New River Estuary (Fig. 1a). Freeman’s Creek is a large (~450,000 m2) polyhaline marsh
located on the Intracoastal Waterway near Brown’s Inlet with a large tidal amplitude (~1m) and a
centralized tidal creek. Smooth cordgrass (Spartina alterniflora) dominates the flora of
Freeman’s Creek. Traps Bay is a pocket-type polyhaline marsh located near the mouth of the
New River Estuary at the New River inlet. Traps Bay is a small marsh (~14,000 m2) dominated
by black needlerush (Juncus roemerianus) with a small, centralized tidal creek and median
(40cm) tidal amplitude. French Creek is a fringing-type marsh (~25,000 m2, 20 to 40m wide)
dominated by black needlerush with no centralized drainage and a small tidal amplitude (20cm).
Unlike the other sites, French Creek is bordered on the upland by a steep slope with a developed
watershed and by oligohaline adjacent waters in the estuary.
Site Development
Each site was developed with a series of piezometers for hydraulic and chemical
sampling, a tide gauge for estuary monitoring, and boardwalks to facilitate access and minimize
marsh disturbance. Piezometers were constructed of 3.175cm diameter PVC with a 40cm long
slotted screen at the base. Upland piezometers were constructed with 180cm long slotted screen.
A 10.96cm diameter auger was used to create the cavity in the marsh in which the piezometers
were set. Shallow piezometers were screened within the rhizoshpere (~0.5m deep) and deep
piezometers were screened below the rhizosphere
10
c
a
b
c
d
b
d
Figure 1. Site map showing Marine Corps Base Camp Lejeune (a), French Creek (b), Traps Bay
(c), and Freeman Creek (d). Images are not to the same scale.
11
(~3m deep). Upland piezometers were screened ~4m below the surface. The slotted screen was
surrounded by a 10.96cm diameter sand pack, capped with grout and bentonite. French Creek
(Fig. 2) consisted of 15 shallow piezometers, 10 deep piezometers, and 2 upland piezometers.
Traps Bay (Fig. 3) consisted of 16 shallow piezometers, 9 deep piezometers, and 2 upland
piezometers.
Freeman’s Creek (Fig. 4) consisted of 12 shallow piezometers, 12 deep
piezometers, and 3 upland piezometers.
Hydraulic Parameters and Water Level Measurements
Hydraulic parameters and water levels were measured in piezometers to determine the
water flux in each marsh (fig. 5). Slug tests were performed to determine hydraulic conductivity
(K) according to the Hvorslev method (Hvorslev 1951). The initial depth to water (DTW) from
the top of the casing (TOC) was taken using a handheld water level meter. An In-Situ Level
Troll100 pressure transducer recording depth at 1-second intervals was then placed in the well at
a known depth from the TOC. Immediately, a volume of water (the slug) was added to the well.
After ~5 minutes, the terminal DTW was read using a handheld water level meter and the
transducer was removed. The data was downloaded and a plot of the ratio of water level (h) to
maximum water level (h0) versus time was generated with the y-axis on a log scale. Hydraulic
conductivity was calculated according to:
K = [r2ln(Le/R)]/[2Let37]
(1)
Where K is hydraulic conductivity, r is the radius of the well casing, R is the radius of the
well screen, Le is the length of the well screen, and t37 is the time required for the water level to
fall to 37% of the initial rise.
Hydraulic head, dh, in selected piezometers was monitored with an In-Situ Level
Troll100 pressure transducer recording water level at fifteen-minute intervals. A separate
12
Figure 2. Aerial photo of French Creek study site. White points represent piezometer clusters.
13
Figure 3. Aerial photo of Traps Bay study site. White dots represent piezometer clusters.
14
Figure 4. Aerial photo of Freeman Creek study site. White dots represent piezometer clusters.
15
Tidal
PN, DIN
DON
Denitrification
N2
Watershed
GW
DIN
Burial
PN
Drainage
DIN / DON,
PN
New
River
Estuary
Figure 5. Conceptual model of salt marsh nitrogen linkages. Tidal recharge supplies particulate
nitrogen (PN), dissolved inorganic nitrogen (DIN), and dissolved organic nitrogen (DON).
Groundwater (GW) inputs supply DIN. Burial removes PN. Denitrification results in gaseous
loss of N2. Porewaters drain to the adjacent estuary resulting in export of DIN, DON, and PN
from the salt marsh system.
16
transducer was left exposed solely to the atmosphere to record barometric pressure at fifteenminute intervals.
The data from the pressure transducers in the wells was corrected for
barometric effects.
A GPS survey was performed to determine the coordinates of each well and elevation
with regard to NAD83 at the TOC for each well. Porewater elevations were calculated as::
hPW = HTOC – DTT + D
(2)
where hPW is the porewater elevation, HTOC is the elevation of the TOC, DTT is the depth to the
transducer, and D is the barometric pressure corrected depth of water recorded by the transducer.
Horizontal distance between water porewater elevation measurements, dx, was determined using
GPS coordinates.
Porewater Transport
Horizontal porewater drainage between any two given piezometers or a piezometer and
the tidal creek was calculated using measured hydraulic head and hydraulic conductivity values
according to Darcy’s Law:
q = -K dh/dx
(3)
where Q is the drainage flux (L m-2 d-1), dh/dx is the hydraulic gradient, and K is the hydraulic
conductivity. Porewater drainage was then brought to the marsh-scale by multiply the marsh
shoreline length by the per meter shoreline drainage.
Chemical sampling
Porewater chemistry was sampled seasonally to be combined with porewater drainage to
calculate solute export fluxes. Porewaters were sampled from all piezometers at all sites for a
range of inorganic and organic analytes.
Sampling occurred seasonally, taking place in
17
February, May, August, and November in 2009 and February, May, and August 2010. All
collections were made at low tide. At each piezometer, water level was recorded using a
handheld water level meter and the piezometer was pumped dry with a peristaltic pump and then
given ~2 minutes to recharge. The porewater was then pumped with a peristaltic pump through
an in-line glass fiber filter and a 0.45μm filter into the appropriate water chemistry vessels for
subsequent analysis of nitrate, nitrite, ammonium, phosphate, dissolve organic carbon/nitrogen,
sulfate, ferrous iron, and hydrogen sulfide. Redox sensitive species SO4, Fe2+, and H2S were
fixed immediately with acetic acid and BaCl2, ferrozine, and diamine reagents in the field,
respectively. DOC/DON samples were H3PO4 acid preserved and refrigerated prior to analysis.
DIN/DIP were frozen and stored until analysis.
Marsh Scale Drainage - Whole Creek Tidal Flood vs. Ebb Flux Studies
At Freeman’s Creek and Traps Bay, duplicate whole creek tidal flux studies were
performed to get a marsh integrated dissolved N export estimate. A tidal flux study could not be
performed at French creek because it is a fringing marsh and lacks a defined tidal creek. Creek
depth profile, current velocity (V), and water chemistry were sampled at 20 to 30 minute
intervals beginning at low tide and finishing at the following low tide. The depth profile was
used to calculate cross-sectional area (A). Discharge (D) was determined using the following
equation:
D = AV
(4)
Samples for water chemistry were collected at the middle of the water column in Traps Bay and
at 4 locations (2 deep, 2 shallow) in Freeman Creek. The same analytes were measured as those
for the seasonal porewater sampling as well as particulate N. Particulate N was sampled by
forcing sample water through a Whatman GFF using a peristaltic pump until the filter was
18
clogged.
The volume of sample filtered was recorded and the filter was sandwiched in
aluminum foil and frozen. Analysis was performed on an Isotope Ratio Mass Spectrometer
(IRMS) elemental analyzer.
Analyte concentrations (specifically DIN and DON) at each
sampling time point were multiplied by the corresponding discharge to generate a nutrient flux
for each sampling time point. Integrating the fluxes over the sampling period generated a marsh
scale flux for the tidal cycle.
RESULTS
Tidal and Watershed Boundary Conditions
Marsh elevation, tidal stage, and water table elevation of the shallow aquifer in the
adjacent watershed served as boundaries for the marsh porewater volume. Marsh elevation
relative to mean high water differed by a factor of 3 between the up-estuary and down-estuary
sites. French Creek, Traps Bay, and Freeman Creek marshes were 14, 18, and 41 cm below mean
high water, respectively. Semi-diurnal tidal patterns observed at all sites were overprinted with
stochastically distributed periods of extended flooding caused by wind-driven high water events
lasting from 1-6 days in duration (Fig. 6). Mean tidal amplitudes (i.e. tidal range) increased with
distance down-estuary from a minimum of 21cm (min = 5.0 cm; max = 51.4cm) at French Creek,
to an intermediate amplitude of 34.3cm (min = 8.4cm; max = 62.1cm) at Traps Bay. The
Freeman Creek marsh exchanged directly with the Intracoastal Waterway (ICW) and was not
influenced by any of the tidal constrictions in the New River Estuary proper. It was the farthest
down-estuary and had the largest mean tidal amplitude of 95.3cm (min = 41.9cm; max =
138.5cm). The combined effects of marsh elevation relative to mean high water and differences
19
a
150
100
50
0
-50
Tidal Elevation (cmNAD83)
-100
Jan-09
Mar-09
Jun-09
Aug-09
Nov-09
Feb-10
Apr-10
b
150
100
50
0
-50
-100
Jan-09
Mar-09
Jun-09
Aug-09
Nov-09
Feb-10
Apr-10
c
150
100
50
0
-50
-100
Jan-09
Mar-09
Jun-09
Aug-09
Nov-09
Feb-10
Apr-10
Figure 6. Daily maximum observed high tide (black) and low tide (grey) for French Creek (a),
Traps Bay (b), and Freeman Creek (c). Creekbank marsh surface elevation is indicated by the
dashed line.
20
in tidal amplitudes among the sites resulted in different flooding durations in the marsh
depending on position within the estuary. While Freeman Creek marsh was flooded
approximately 40% of the time during the yearlong monitoring period, Traps Bay marsh was
flooded 38%, and French Creek was flooded 62% of the time.
Upland water table elevations adjacent to all marshes were characterized by seasonal
fluctuations of approximately 50 cm, and were interspersed with short duration excursions of
almost equal magnitude that were attributable to tidal effects and/or precipitation events (Figs. 7,
8). Seasonal drawdown of the water table was seen in all watersheds during summer 2009 and
summer 2010. The absolute elevation of the water table was largely a function of the distance
between the location of the upland well and the marsh-upland edge (French Creek=15m; Traps
Bay=15m; Freeman Creek=100m). The slopes of the water table towards the marsh were similar
between French Creek and Traps Bay, but a factor of 4 smaller at the lower topographical
gradient and larger Freeman Creek site. A dampened tidal effect and rapid response of the water
table to precipitation events were encountered in the Freeman Creek and Traps Bay uplands (Fig.
7,8). A tidal signature in the Traps Bay upland was only evident during periods of low water
table. Short-duration fluctuations in the French Creek water table were small in comparison to
Traps Bay and Freeman Creek, were wholly tidally driven, and showed little to no response to
precipitation events (Fig. 7a).
Hydraulic Properties of the Rhizosphere
Horizontal hydraulic conductivity (Kh) of the rhizosphere (0-50cm depth) determined by
in-situ slug tests were on the order of that typically found in fine sands (Table 1). Kh averaged
2.3 ± 0.5 × 10-3 cm s-1 for French Creek, 4.1 ± 0.7 × 10-3 cm s-1 for Traps Bay, and 1.7 ± 0.4 ×
10-3 cm s-1 for Freeman Creek. Mean Kh for the interior of French Creek was twice
21
a
150
100
Watershed Water Table Elevation (cm NAD83)
50
0
-50
Apr-09
Jul-09
Sep-09
Dec-09
Mar-10
May-10
Aug-10
b
Oct-10
Jul-09
Sep-09
Dec-09
Mar-10
May-10
Aug-10
c
Oct-10
Jul-09
Sep-09
Dec-09
Mar-10
May-10
Aug-10
Oct-10
150
100
50
0
-50
Apr-09
150
100
50
0
-50
Apr-09
Figure 7. Upland water table elevations for French Creek (a), Traps Bay (b), and Freeman Creek
(c). Freeman Creek and Traps Bay showed responses to precipitation events while French Creek
predominantly responded to tidal elevation.
22
a
150
100
3
50
2
0
1
-50
12/6
12/11
12/16
12/21
0
12/31
12/26
b
150
100
4
3
50
2
0
-50
1
-100
0
12/1
12/6
12/11
12/16
12/21
12/26
12/31
c
150
100
4
3
50
2
0
-50
1
-100
0
12/1
Precipitation (cm)
Water Elevation (cm NAD83)
-100
12/1
4
12/6
12/11
12/16
12/21
12/26
12/31
Figure 8. Short duration fluctuations in upland water table elevation (grey), tidal elevation
(black), and precipitation (bar) for French Creek (a), Traps Bay (b), and Freeman Creek (c).
23
that of mean creekside Kh. With the exception of the most upstream transects at Traps Bay
dominated by Typha spp, mean interior Kh was 2-3 times greater in the interior of the marshes
than in the creekbanks for all sites (Table 1). There were no significant inter-marsh differences
between interior Kh or creekbank Kh values. Vertical hydraulic conductivity (Kv) values
determined from falling head permeameter measurements on rhizosphere cores were on average
10% of Kh values for the same marsh locations.
The horizontal hydraulic gradient was considered separately for the creekbank and
interior zones of the marsh. The creekbank gradients were calculated from the difference in
hydraulic head between the creekbank piezometer and tidal stage, and the marsh interior gradient
was calculated from the difference in hydraulic head measured at the upland-marsh border and
the creekbank piezometers. The daily maximum hydraulic gradient towards open water (i.e.
drainage) in the creekbank exceeded the interior gradient by a factor of 10-100. The creekbank
gradient was dynamic and exhibited hourly variation throughout the tidal cycle, while
fluctuations on hourly to daily scales in the interior gradient were barely detectable at all sites
(Fig. 9). The gradient from the creekbank to open water was maximized at low tide for all sites.
Consistent with the different marsh elevations relative to mean high water and the different tidal
amplitudes, the low tide creekbank hydraulic gradients among sites differed by a factor of 10
(Fig. 9) with a maximum and minimum occurring at Freeman Creek and French Creek,
respectively. For all marshes, as the tide rose the creekbank gradient decreased until the tidal
elevation matched or exceeded the porewater elevation (Fig. 9). When tidal elevation was
greater than porewater elevation, either no gradient existed, or a negative gradient was created
(Fig. 9 a,b), indicating periods of recharge. Mean positive creekbank gradient increased with
distance down-estuary from French Creek (0.021) and Traps Bay (0.025) to Freeman Creek
24
a
40
20
0
-20
-40
-60
-80
9:36
b
40
20
0
-20
-40
-60
-80
19:12
c
40
20
0
-20
-40
-60
-80
0.35
0.25
0.15
0.05
-0.05
0:00
4:48
9:36
14:24
19:12
0:00
4:48
0.45
dh/dx
0.35
0.25
0.15
0.05
-0.05
14:24
19:12
0:00
4:48
9:36
14:24
0.45
0.35
0.25
0.15
0.05
-0.05
0:00
4:48
9:36
14:24
19:12
0:00
4:48
Tidal Elevation (cm NAD83)
0.45
9:36
Figure 9. Representative daily fluctuation of tidal elevation (grey), creekbank gradient (triangle),
and interior gradient (circle) for French Creek (a), Traps Bay(b), and Freeman Creek (c). Dh
represents the difference in hydraulic head between either the tide gauge and creekbank
piezometer (creekbank gradient), or the upland-marsh border piezometer and the creekbank
piezometer (interior gradient). Dx represents the distance between the points where dh is
determined. A positive gradient indicates a slope from the marsh towards open water.
25
Table 1. Hydraulic conductivity (Kh) of Traps Bay, French Creek, and Freeman Creek.
K h x10-3 cm s-1
Well location
Transect
Creekside Interior
Traps 1 East
6.32
3.72
Traps 1 West
9.03
9.03
Traps 3 East
7.90
1.86
Traps 3 West
4.10
Traps 5 East
1.44
7.90
Traps 5 West
2.60
1.58
Traps 7 East
0.84
1.34
Traps 7 West
1.41
2.04
French 1
0.93
5.26
French 3
0.67
0.79
French 7
0.73
3.72
French 9
1.32
4.20
Freeman 1
1.47
1.19
Freeman 3
0.97
0.63
Freeman 5
1.40
5.50
26
(0.103). Creekbank gradient was negative in Freeman creek for 6% of the monitoring period
compared to 43% in French Creek and 49% in Traps Bay. Mean negative creekbank gradient
was greatest in French Creek (-0.019) and decreased with distance down-estuary to Traps Bay (0.015) and was smallest in Freeman Creek (-0.012).
Interior drainage gradients were on the order of 20% of the creekbank gradient in French
Creek, to less than 2% of the creekbank gradient in Freeman Creek. Mean positive interior
drainage gradient decreased with distance down-estuary from 0.004 in French Creek to 0.003 in
Traps Bay and 0.001 in Freeman Creek. Mean negative interior drainage was smallest in
Freeman Creek (-0.001, 29% of sampling period) and increased with distance up-estuary to
Traps Bay (-0.002, 44% of sampling period) and French Creek (-0.004, 34% of sampling
period).
Negative vertical gradient (dh/dz) indicated vertical recharge during flooded conditions.
As the marsh flooded and tidal elevation rose above porewater elevation (Fig. 10), vertical
gradient became negative (Fig. 11). Mean negative vertical gradient was -0.14 in French Creek,
-0.15 in Traps Bay, and -0.01 in Freeman Creek.
Porewater Drainage Drivers
Using all transducer records aggregated over an annual period, daily drainage of the
creekbank to adjacent waters and from the interior of the marsh to the creekbank were calculated
using Eq. 3. Overall magnitudes, temporal patterns of change, and responses of each of these
drainage fluxes in response to changes in tidal and upland water table elevation were assessed for
each site. Drainage in the creekbank was greater than drainage from the marsh interior for all
sites.
27
a
40
20
0
Water Elevation (cm NAD83)
-20
-40
b
40
20
0
-20
-40
c
40
20
0
-20
-40
00:00
12:00
00:00
Figure 10. Creekbank porewater elevation (triangles) and tidal elevation (black line) through a
daily tidal series for French Creek (a), Traps Bay (b), and Freeman Creek (c). Once the marsh is
flooded, tidal elevation is greater than porewater elevation, creating a negative gradient. Shaded
grey areas represent times of vertical recharge.
28
a
-0.1
20
-0.2
0
-0.3
-20
-0.4
-40
-0.5
-60
0:00
4:48
9:36
14:24
19:12
0:00
4:48
9:36
b
0
dh/dz
40
40
-0.1
20
-0.2
0
-0.3
-20
-0.4
-40
-0.5
-60
14:24
19:12
0:00
4:48
9:36
14:24
19:12
c
0
40
-0.1
20
-0.2
0
-0.3
-20
-0.4
-40
-0.5
-60
19:12
0:00
4:48
9:36
14:24
19:12
0:00
Tidal Elevation (cm NAD83)
0
4:48
Figure 11. Vertical gradient for the creekbank (triangles) and marsh interior (circles) for French
Creek (a), Traps Bay (b), and Freeman Creek (c) during flooded periods of a representative tidal
cycle. Tidal elevation is indicated by the black line. Negative gradients indicate vertical
recharge of the marsh.
29
Creekbank Drainage and Tides
For all creekbanks at all sites, a two-phase response was found between porewater
drainage and tidal elevation (Figs. 12-14). Drainage increased as a function of decreasing tidal
elevation below the marsh surface. Except for the two most up-gradient transects at Traps Bay
whose drainage was heavily influenced by groundwater discharge, the largest increase in
drainage per unit drop in tidal height was found in Freeman Creek, followed by Traps Bay, and
French Creek. For every cm tidal elevation fell below the marsh surface, Freeman Creek, Traps
Bay, and French Creek drainage fluxes increased by 3.5, 2.0, and 0.85 L m-2 d-1. The higher
drainage fluxes per unit tidal change also showed less variability in the range of drainage for a
given tidal elevation. This phenomenon was evidenced by the higher correlation coefficients
relating tidal stage to drainage for Freeman Creek and Traps Bay relative to French Creek. At
tidal elevations above the marsh surface, the drainage flux (either a small residual flux towards
open water, or a negative recharge flux) was largely independent of further changes in tidal
height so long as the marsh was flooded (Figs. 12-14; gray regions).
Creekbank Drainage and Upland Water Table
The relationship between creekbank drainage and upland water table slope was examined
for the lowest daily tidal elevation values. These periods coincided with maximum drainage of
marsh porewater, and were assumed to be most indicative of water table effects not confounded
by tidal effects. French Creek had a significant daily maximum creekbank drainage response to
increased water table slope in all transects (Fig. 15). Traps Bay showed a significant daily
maximum creekbank drainage response to increased upland water table slope for all
30
a
100
y = -0.639x + 11.2
2
R = 0.132
p<0.001
50
0
Creekbank Drainage (L m-2 day-1)
-50
-20
-10
0
10
20
100
y = -1.23x + 20.8
50
R = 0.159
p<0.001
30
40
50
60
b
2
0
-50
-20
-10
0
10
20
100
y = -0.719x - 0.596
50
R = 0.171
p<0.001
30
40
50
60
c
2
0
-50
-20
-10
0
10
20
30
40
50
60
Tide Elevation (cm NAD83)
Figure 12. Creekbank drainage versus tidal elevation for French Creek transect 3 (a), transect 5
(b), and transect 7 (c). A two-phase response was seen with increased drainage at elevations
below the marsh surface (black ) and recharge for tidal elevations greater than the marsh surface
(grey). Grey points greater than zero occur as the marsh is draining on the falling tide and grey
values less than zero occur as the marsh is flooding on the rising tide.
31
a
800
600
400
200
0
-200
-400
y = -9.85x + 41.7
2
R = 0.399
p<0.001
Creekbank Drainage (L m-2 day-1)
-40
-20
0
20
40
60
80
100
b
800
600
400
200
0
-200
-400
y = -14.0x + 34.1
2
R = 0.284
p<0.001
-40
-20
0
20
40
60
80
100
c
800
600
400
200
0
-200
-400
y = -2.63x + 3.58
2
R = 0.655
p<0.001
-40
-20
0
20
40
60
80
100
d
800
600
400
200
0
-200
-400
y = -1.56x + 4.35
2
R = 0.561
p<0.001
-40
-20
0
20
40
60
80
100
Tide Elevation (cm NAD83)
Figure 13. Creekbank drainage versus tidal elevation for Traps Bay transect 1 (most upstream)
(a), transect 3 (b), transect 5 (c), and transect 7 (most downstream) (d). A two-phase response
was seen with increased drainage at elevations below the marsh surface (black ) and recharge for
tidal elevations greater than the marsh surface (grey). Grey points greater than zero occur as the
marsh is draining on the falling tide and grey values less than zero occur as the marsh is flooding
on the rising tide.
32
400
a
y = -4.17x + 83.5
300
2
R = 0.932
p<0.001
200
100
0
Creekbank Drainage (L m-2 day-1)
-100
-100
-50
0
50
100
b
400
y = -2.60x + 25.4
300
2
R = 0.912
p<0.001
200
100
0
-100
-100
-50
0
50
100
c
400
y = -3.97x + 46.6
300
2
R = 0.971
p<0.001
200
100
0
-100
-100
-50
0
50
100
Tide Elevation (cm NAD83)
Figure 14. Creekbank drainage versus tidal elevation for Freeman Creek transect 1 (a), transect 3
(b), and transect 5 (c). A two-phase response was seen with increased drainage at elevations
below the marsh surface (black ) and recharge for tidal elevations greater than the marsh surface
(grey). Grey points greater than zero occur as the marsh is draining on the falling tide and grey
values less than zero occur as the marsh is flooding on the rising tide.
33
transects (Fig.16). Freeman Creek showed a significant daily maximum creekbank drainage
response to increased upland water table slope for all transects (Fig. 17) The magnitude of
response was greatest in Traps Bay, less in Freeman Creek, and least in French Creek.
Interior Drainage and Tides
Drainage from the interior of the marsh to the creekbank was influenced by tidal elevations in all
marshes (Figs. 18-20). In general, the magnitude of interior drainage was less than that of
creekbank drainage. As a percentage of creekbank drainage, low tide interior drainage rates
were 10-50% for French Creek, 5-40% for Traps Bay (lower transects), and less than 2% for
Freeman Creek. The response pattern of interior drainage to tidal elevation varied by marsh.
The response in French Creek (Fig. 18) was a two-phase response (marsh flooded or not) similar
to that seen with creekbank drainage (Fig. 12). However, Traps Bay and Freeman Creek were
characterized by a three-phase response that depended on whether the marsh was flooded, and if
so, by how much. For Traps Bay, when the marsh initially flooded from 10cm tidal elevation to
20cm tidal elevation, interior drainage decreased with rising water and in some cases flipped to
recharge (Fig. 19c,d). Only after the marsh was flooded (tidal elevation> 20cm) did the interior
of the marsh recharge (negative interior drainage) whereby the drainage/recharge was unaffected
by more overlying water. At tidal elevations less than the marsh surface elevation, a slight
increase in drainage was seen with decreasing tidal elevation. A similar pattern was observed in
Freeman Creek where the critical depth of overlying flooded water ranged between 10 and 50 cm
depending on location in the marsh (Fig. 20).
34
a
100
50
y = 37400x - 9.64
Daily Maximum Creekbank Drainage (L m-2 d-1)
0
2
R = 0.413
p<0.001
-50
0
0.0005
0.001
0.0015
0.002
b
100
50
y = 54000x - 4.03
0
2
R = 0.49
p<0.001
-50
0
0.0005
0.001
0.0015
0.002
c
100
50
y = 21200x - 4.73
0
2
R = 0.129
p<0.001
-50
0
0.0005
0.001
0.0015
0.002
Water Table Slope
Figure 15. Daily maximum creekbank drainage response to upland water table slope for French
Creek transect 3 (a), transect 5 (b), and transect 7 (c).
35
600
a
y = 1E+06x - 430
400
R2 = 0.685
p<0.001
200
0
-200
0
Daily Maximum Creekbank Drainage (L m-2 d-1)
600
0.0002
0.0004
0.0006
0.0008
0.001
b
y = 939000x - 158
400
2
R = 0.362
p<0.001
200
0
-200
0
600
0.0002
0.0004
0.0006
0.0008
0.001
c
y = 200000x - 42.4
400
2
R = 0.558
p<0.001
200
0
-200
0
600
0.0002
0.0004
0.0006
0.0008
0.001
d
y = 125000x - 35.7
400
2
R = 0.615
p<0.001
200
0
-200
0
0.0002
0.0004
0.0006
0.0008
0.001
Water Table Slope
Figure 16. Daily maximum creekbank drainage response to upland water table slope for Traps
Bay transect 1 (a), transect 3 (b), transect 5 (c), and transect 7 (d).
36
a
400
300
200 y = 112000x + 216
2
R = 0.401
p<0.001
Daily Maximum Creekbank Drainage (L m-2 d-1)
100
0
0
0.0005
0.001
0.0015
0.002
b
400
300
200 y = 48600x + 109
2
R = 0.160
p<0.001
100
0
0
0.0005
0.001
0.0015
0.002
c
400
300
200y = 64000x + 195
2
R = 0.111
p<0.001
100
0
0
0.0005
0.001
0.0015
0.002
Water Table Slope
Figure 17. Daily maximum creekbank drainage response to upland water table slope for Freeman
Creek transect 1 (a), transect 3 (b), and transect 5(c).
37
a
20
10
y = -0.135x + 1.76
R² = 0.208
p<0.001
0
-10
-20
Interior Drainage (L m-2 d -1 )
-20
-10
0
10
20
30
40
50
60
b
20
10
0
-10
-20
y = -0.229x + 5.26
R² = 0.357
p<0.001
-20
-10
0
10
20
30
40
50
60
c
20
10
0
-10
-20
y = -0.287x + 11.9
R² = 0.258
p<0.001
-20
-10
0
10
20
30
40
50
60
Tidal Elevation (cm NAD83)
Figure 18. Interior marsh drainage response to tidal elevation for French Creek transect 3 (a),
transect 5(b), and transect 7 (c). A two-phase response was seen for tidal elevations below
interior marsh surface (black ) and above interior marsh surface (grey).
38
Figure 19. Interior marsh drainage response to tidal elevation for Traps Bay transect 1 (a),
transect 3 (b), transect 5 (c), and transect 7 (d). A three-phase response was seen for tidal
elevations below creekbank marsh surface (black ), between creekbank marsh surface and
interior marsh surface (light grey), and above interior marsh surface (dark grey).
39
a
10
y = -0.126x + 3.08
5
R = 0.201
p<0.001
2
0
-5
Interior Drainage (L m-2 d-1)
-10
-100
-50
0
50
10
y = -0.142x + 1.72
5
R = 0.042
p<0.001
100
b
2
0
-5
-10
-100
-50
0
50
10
y = -0.088x + 5.05
5
R = 0.342
p<0.001
100
c
2
0
-5
-10
-100
-50
0
50
100
Tidal Elevation (cm NAD83)
Figure 20. Interior marsh drainage response to tidal elevation for Freeman Creek transect 1 (a),
transect 3 (b), and transect 5 (c). A three-phase response was seen for tidal elevations below
creekbank marsh surface (black ), between creekbank marsh surface and interior marsh surface
(light grey), and above interior marsh surface (dark grey).
40
a
20
10
0
Daily Maximum Interior Drainage (L m-2 d-1)
-10
-20
0
0.0002
0.0004
0.0006
0.0008
0.001
0.0012
20
0.0014
b
10
0
-10
p<0.1
-20
0
0.0002
0.0004
0.0006
0.0008
0.001
0.0012
0.0014
c
20
10
y = 7360x + 10.2
0
2
R = 0.450
p<0.001
-10
-20
0
0.0002
0.0004
0.0006
0.0008
0.001
0.0012
0.0014
Water Table Slope
Figure 21. Daily maximum interior drainage response to upland water table slope for French
Creek transect 3 (a), transect 5 (b), and transect 7(c).
41
a
60
40
20
y = 25200x - 0.479
0
2
R = 0.672
p<0.001
-20
-40
0
0.0005
0.001
0.0015
0.002
0.0025
0.003
b
Daily Maximum Interior Drainage (L m-2 d-1)
60
40
20
y = 9420x + 9.52
0
2
R = 0.135
p<0.005
-20
-40
0
0.0005
0.001
0.0015
0.002
0.0025
0.003
c
60
40
20
y = 4770x + 0.192
0
2
R = 0.184
p<0.001
-20
-40
0
0.0005
0.001
0.0015
0.002
0.0025
60
0.003
d
40
20
0
-20
p>0.15
-40
0
0.0005
0.001
0.0015
0.002
0.0025
0.003
Water Table Slope
Figure 22. Daily maximum interior drainage response to upland water table slope for Traps Bay
transect 1 (a), transect 3 (b), transect 5 (c). and transect 7(d).
42
a
10
Daily Maximum Interior Drainage (L m-2 d-1)
5
0
y = 1030x + 1.54
-5
R2 = 0.054
p<0.05
-10
0
0.0005
0.001
0.0015
0.002
b
10
5
0
y = 4480x - 0.417
-5
R = 0.284
p<0.001
2
-10
0
0.0005
0.001
0.0015
0.002
c
10
5
0
y = 3570x + 2.25
-5
R = 0.438
p<0.001
2
-10
0
0.0005
0.001
0.0015
0.002
Water Table Slope
Figure 23. Daily maximum interior drainage response to upland water table slope for Freeman
Creek transect 1 (a), transect 3 (b), and transect 5(c).
43
Interior Drainage and Upland Water Table
Seven of the ten transects monitored among all the marshes showed a positive correlation
between maximum daily interior drainage and changes in water table slope (Figs. 21-23). The
three transects that did not show a correlation between changes in water table elevation and
interior drainage (two at French Creek and one in Traps Bay) possessed the lowest overall
magnitudes of interior drainage at their respective sites. Of the significant transects, French
Creek showed the greatest per unit increase in interior drainage in response to increased upland
slope followed by Traps Bay and Freeman Creek.
Porewater Drainage Flux – Summary
While porewater drainage through the creekbank (through which interior drainage must
pass) within a site was related most strongly to tidal elevation (Figs. 12-14, 18-20), marsh
drainage trended with daily tidal amplitude at the estuarine scale (Fig. 24).
Greater tidal
amplitude yielded greater maximum creekbank drainage.
Monthly-averaged, Darcy-derived estimates of creekbank drainage were used to scale up
time series drainage fluxes to the marsh-scale. Per-square meter fluxes (L m-2 d-1) were scaled to
the cross-sectional face of the creekbank for each marsh based upon the height of the creekbank
relative to mean low water and mean tidal amplitude at the site such that the resultant flux was
given units of L m shoreline-1 day-1. The monthly estimates of drainage at each site varied
several-fold over an annual period, but showed no seasonal patterns. Monthly average porewater
drainage in Freeman Creek (Fig. 25c, 44.4-123.6 L m shoreline-1 d-1) was 20 times greater than
44
Daily Maximum Creekbank Drainage (L m-2 d -1 )
250
200
150
100
50
0
0
50
100
Daily Tidal Amplitude (cm)
150
Figure 24. All sites daily maximum drainage versus daily tidal amplitude. French Creek (open
circles), Traps Bay (black circles), and Freeman Creek (triangles).
45
Monthly Mean Creekbank Drainage (L m shoreline-1 d-1)
1000
100
10
1
0.1
0.01
0.001
J F M A M J J A S O N D J F M A M J
0 1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19
Freeman
French
Traps
Figure 25. Mean monthly drainage corrected for cross sectional drainage area for French Creek
(solid triangles), Traps Bay (open boxes), and Freeman Creek (solid diamonds). Error bars
represent standard error.
46
in Traps Bay (Fig. 25b, 0.5-5.9 L m shoreline-1 d-1) and Traps Bay was 5 times greater than in
French Creek (Fig. 25a, 0.01-0.9 L m shoreline-1 d-1) each month. Greater drainage generally
occurred during periods of lower mean tidal elevation for all sites, such that longer periods of
tidal inundation, such as those associated with wind-driven flooding, equated with little to no
drainage. Traps Bay monthly mean drainage per day differed by an order of magnitude between
upstream (groundwater influenced) and downstream transects while drainage was more
consistent between transects at the other sites.
The percentage creekbank drainage that was supplied by interior drainage to the
creekbank increased with distance up-estuary. Freeman Creek transect 1 (0.8±0.1%), transect 3
(1.1±0.2%) and transect 5 (3.6±0.8%) had the lowest percent creekbank drainage supplied by
interior drainage.
Traps Bay transect 1 (8.2±2.3%), transect 3 (6.9±2.4%), and transect 7
(4.9±0.8%) had the second lowest percent contribution of interior drainage to total creekbank
drainage while the interior of transect 5 contributed 42.4±8.2% of the total creekbank drainage.
French Creek interior drainage contribution to creekbank drainage in transect 3 (15.6±4.5%) and
transect 5 (42.2±22.5%) was higher than in Freeman Creek and Traps Bay. French Creek
interior of transect 7 contributed 154.5±48.6% to the creekbank drainage, suggesting an alternate
cross transect flowpath existed.
Porewater Nitrogen Flux from Marsh to Open Water
The total dissolved nitrogen (DIN + DON) flux attributable to porewater drainage was
derived from the monthly aggregated porewater drainage fluxes and seasonally measured
porewater TDN values. Total dissolved nitrogen varied within the marsh and the highest
concentrations were found between the creekbank and the interior. Mean creekbank TDN
concentrations were greatest during the winter (Table 2). Mean TDN concentrations were
47
Table 2. Quarterly porewater concentrations of total dissolved nitrogen (TDN) based on location within each marsh
Feb-09
May-09
Aug-09
Nov-09
Feb-10
May-10
French
Creekside
71 ± 11
155 ± 41
210 ± 60
218 ± 69
243 ± 87
184 ± 54
Mean porewater TDN concetration (µM)
Traps
Traps
Freeman
French
Creekside
Interior
Creekside
Interior
43 ± 17
68 ± 12
55 ± 16
40 ± 11
151 ± 99
65 ± 12
95 ± 41
28 ± 4
126 ± 44
106 ± 23
86 ± 28
69 ± 9
130 ± 51
175 ± 40
110 ± 24
144 ± 46
204 ± 77
122 ± 19
99 ± 15
128 ± 38
186 ± 86
124 ± 20
107 ± 19
58 ± 14
48
Freeman
Interior
57 ± 20
32 ± 9
44 ± 25
59 ± 28
108 ± 9
53 ± 17
greatest in the marshes that had the least drainage with annual mean TDN concentration greatest
in French Creek (158 ± 15µM), followed by Traps Bay (103 ± 8µM), and lowest in Freeman’s
Creek (85 ± 9). For all porewaters, NO2+NO3 was negligible (generally <1uM) and TDN was
comprised of approximately 50-70% DON and 30-50% NH4+.
When coupled with the
creekbank porewater drainage estimates (Fig. 25), the advective fluxes of nitrogen showed a
seasonal pattern in all three sites with a peak at the end of summer and a second more prolonged
peak at the beginning of winter (Fig. 26). French Creek showed the least seasonal variation (.021.7 g N-TDN day-1) and the least drainage, followed by Traps (0.7-9.2 g N-TDN day-1), and
Freeman Creek showed the greatest temporal variation in N flux (44-120 g N-TDN day-1)
accompanied by the greatest drainage (Fig. 27). Despite higher TDN concentrations in the upestuary marshes, the inter-marsh differences in N flux was almost solely attributable to
differences in porewater drainage rates.
Comparison of Darcy N Flux with Tidal N Exchange Estimates
Estimates of whole marsh N import/export were also obtained by comparing flood and
ebb N fluxes during a tidal cycle in the Traps Bay and Freeman Creek sites; both of which
flooded and drained to well- defined tidal creeks. Traps Bay was sampled where the creek meets
open water and the creek width is <5m. Traps Bay was sampled September 10, 2009 and May 3,
2010. Freeman Creek was sampled approximately 0.5 km upstream from the mouth of the creek
where creek width was ~35m. Freeman Creek was sampled September 11, 2009 and May 4,
2010.
For Traps Bay, the September 2009 sampling had an initial low tide at 07:45 (Fig. 28a,b).
High tide was at 13:15 and final low tide was at 20:30. Tidal amplitude for the sampling was
31.6cm. For the sampling period, the marsh showed a net import of 60.7g N-TDN, a net import
49
of 4.5g N-PON, a net import of 500 g N-NH4+, and a net export of 435g N-DON. The May 2010
sampling occurred during a period of extended low water (Fig. 28c,d). High tide occurred at
12:15 with a tidal amplitude of 16cm. During the sampling period, the net marsh export of 83.8g
N-NH4+, 22g N-NO3-, 3310g N-TDN, 3210g N-DON, and 367g N-PON. The large TDN and
DON exports resulted from an increase in DON beginning around 14:00.
For Freeman Creek, The September 2009 sampling initial low tide was at 06:00 (Fig.
29a,b). High tide was at 13:00 and final low tide was at 18:45. Tidal amplitude was 94cm on
the rising tide and 74cm on the falling asymmetrical tide. During the sampling period, the marsh
showed a net import of 3650g N-NH4+, 2220g N-NO3-, 30200g N-TDN, 24200g N-DON, and
14700g N-PON. The May 2010 sampling occurred during a period of extended low water.
Initial low tide was at 07:15 with high tide occurring at 11:30 and failing to crest the marsh
platform (Fig. 29c,d). Final low tide was at 18:30. During the sampling period, the marsh
posted a net export of 250g N-NH4+, and 108g N-PON. The marsh posted a net import of 46g NNO3-, 74800g N-TDN, and 74900g N-DON, There was a major pulse of high concentration
DON at the beginning of the sampling that dropped to ~10% of initial concentration around
15:00 which was responsible for the large influx of DON and TDN.
50
Monthly Mean Creekbank Drainage (L m shoreline-1 d -1 )
1000
100
10
1
0.1
0.01
0.001
J F M A M J J A S O N D J F M A M J
0 1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19
Freeman
French
Traps
Figure 26. Monthly average nitrogen drainage for French Creek, Traps Bay, and Freeman Creek.
Freeman Creek drains two orders of magnitude more N per day than Traps Bay and French
Creek.
51
1000
mg N-TDN m shoreline-1 d -1
100
10
1
0.1
0.01
0
J F M A M J J A S O N D J F M A M J
1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19
French
Traps
Freeman
Figure 27. Monthly average Darcy-derived nitrogen flux for French Creek, Traps Bay, and
Freeman Creek. Freeman Creek drains two orders of magnitude more N per day than Traps Bay
and French Creek. Rates were scale up to the whole marsh using shoreline length estimates of
1300m for French Creek, 900m for Traps Bay, and 9600m for Freeman Creek. Open diamonds
represent Freeman Creek tidal flux estimates out of the marsh. Closed box represents Traps Bay
tidal flux out of the marsh. Plus sign represents Traps Bay tidal flux into the marsh.
52
300
500
TDN (µM)
300
700
500
300
200
100
150
100
50
100
150
-100
100
-300
50
0
7:40
-500
10:24
12:48
15:12
17:36
20:00
b
50
40
30
30
20
20
10
10
0
0
-10
-10
-20
-20
10:24
12:48
15:12
17:36
-100
-300
10:33
13:26
-30
20:00
7:40
-500
19:12
16:19
d
50
40
-30
8:00
c
300
200
0
8:00
350
250
250
Tidal Elevation (cm NAD83)
700
Discharge (L s-1)
a
350
10:33
13:26
16:19
19:12
Figure 28. Traps Bay discharge L s-1 (gray area) (a,c), TDN concentration (black line) (a,c), and tidal signature (b,d). Discharge out of
the marsh is negative, discharge into the marsh is positive. Tidal flux studies from September 10, 2009 (a,b) and May 3, 2010 (c,d)
are shown.
53
300
TDN (µM)
250
c
350
20000
300
10000
250
200
10000
200
0
150
150
100
0
7:55
10:19
12:43
15:07
17:31
-10000
100
-20000
50
60
40
40
20
20
0
0
-20
-20
-40
-40
-60
-60
12:43
15:07
17:31
-20000
-30000
10:48
13:12
15:36
18:00
d
80
60
10:19
-10000
0
8:24
-30000
19:55
b
80
-80
7:55
30000
20000
0
50
Tidal Elevation (cm NAD83)
30000
Discharge (L s-1)
a
350
-80
19:55
8:24
10:48
13:12
15:36
18:00
Figure 29. Freeman Creek discharge L s-1 (gray area) (a,c), TDN concentration (black line) (a,c), and tidal signature (b,d). Discharge
out of the marsh is negative, discharge into the marsh is positive. Tidal flux studies from September 11, 2009 (a,b) and May 4, 2010
(c,d) are shown.
54
DISCUSSION
Water and N Movement
Water Movement-Creekbank
Porewater draining through the creekbank to the adjacent open water was derived from
two main sources: 1) tidal water recharged during high tide; and 2) porewater draining from the
interior of the marsh into the creekbank. At low tide, porewater elevation in the creekbank was
higher than tidal elevation, indicating horizontal drainage toward the discharge interface between
the creekbank and open water (Fig. 9). As the tide rose, the hydraulic gradient from creekbank
to open water decreased until tidal elevation reached the porewater elevation, at which point the
gradient reversed.
The horizontal gradient now moved from the open water toward the
creekbank began increasing. During this period of rising but non-flooding tide, there was a brief
period of weak horizontal recharge.
Once the marsh was overtopped the overlying water
recharged the creekbank vertically, as evidenced by a vertical hydraulic gradient resulting in
downward flow between flooded water and marsh porewater. The recharge during flooding was
rapid. In Freeman Creek, recharge occurred within 1 hour following marsh flooding and was
independent of the depth of flooding water. In French Creek and Traps Bay the recharge was
more prolonged (extending 3-6 hours) and its magnitude was dependent on the depth of the
water overlying the marsh. The falling tide elicited no porewater drainage until the marsh
surface was no longer flooded, at which point porewater drainage increased as tidal elevation
decreased. In general this pattern of dominant vertical tidal recharge followed by horizontal
drainage is consistent with that reported by Harvey et al. (1987) and Gardner (2005).
55
Exceptions to this pattern of tidal recharge and drainage occurred during periods of
extended high water when the marsh did not completely drain and when the tide was
transitioning from a period of prolonged flooding to a more regular pattern of wetting and
drying. In both of these cases, the porewater elevation remained higher than tidal elevation
during the flooded low tides, indicating vertical discharge up through the marsh surface. Vertical
discharge occurred on less than 10% of the tides in French Creek, less than 5% of the tides in
Traps Bay, and almost less than 1% of the tides in Freeman Creek. These instances of vertical
discharge occurred on higher tides when the upland water table was pushed to a higher elevation.
The flux of porewater up through the marsh was supported by the existence of a vertical head
gradient from the deep piezometers to the shallow piezometers. These anomalous periods of
vertical discharge seem to occur more with smaller tidal amplitudes and likely never occurred in
Freeman Creek because the marsh was able to fully drain on every tidal cycle.
Water Movement-Interior
The contribution of porewater draining from the interior of the marsh relative to tidal
recharge of the creekbank ranged from about 50% in French Creek to less than 5% in Freeman
Creek. The interior drainage responded to both tidal elevation and water table elevation. The
interior followed similar patterns of drainage and recharge with the tides as described for the
creekbank. However there was a stronger connection between interior drainage and the upland
than between the creekbank and the upland. This connection was evidenced by the strong
correlation between interior drainage and upland water table elevation when drainage was largest
(low tide; Figs. 21-23). Similar to the two-phase drainage response to tidal elevation (Figs. 1214) found in the creekbank of all sites, the interior drainage of French Creek showed a two-phase
response to tidal elevation whereby drainage increased at tidal elevations below the marsh
56
surface and remained near zero at tidal elevations above the marsh surface, indicating that the
type of response is consistent through the entire marsh of French Creek.
The importance of interior drainage flux relative to tidal recharge was directly related to
water table slope, but inversely related to tidal amplitude and width of the marsh. Interior
drainage contributed less than 5% of the porewater in Freeman Creek, which had a marsh width
that was four times greater and a water table slope four times less than French Creek and Traps
Bay. The interior drainage contribution to the creekbank was greatest in the narrower marshes
with greater interior hydraulic gradients (French Creek and Traps Bay). This pattern was similar
to observations of Tobias et al. (2001b) that showed that watershed controlled groundwater
discharge in a fringing marsh in Virginia facilitated export of solutes from marsh porewater to
estuary. The larger contribution of interior drainage to maintaining creekbank volume in Traps
and French suggests that export of solutes via drainage through the creekbank in low tidal
amplitude narrow marsh morphologies are more closely dependent on subsurface advection
through the rest of the marsh, while creekbank drainage of expansive high tidal amplitude
marshes is largely a near-bank process uncoupled from the marsh interior.
The combined effects of drainage and recharge (tidal and from the marsh interior)
permitted estimation of porewater residence times in the creekbanks of each marsh. Assuming
steady-state creekbank porewater elevations on monthly timescales and a creekbank distance of
3m normal to the shoreline per 1m of shoreline, the porewater volume of the creekbank was 0.18
m3 in French Creek, 0.3 m3 in Traps Bay, and 0.9 m3 in Freeman Creek. The calculated
porewater residence times for French Creek, Traps Bay, and Freeman Creek marshes are on the
order of 360, 90, and 12 days, respectively. Even the shortest of the residence times (Freeman
Creek) is much longer than the timescale of internal N-cycling reactions (Tobias and Neubauer
57
2009). These residence times are directly related to tidal elevation within a site and amplitude
between sites (Fig. 6) but overall the inventory of solutes (e.g. TDN) is overwhelmingly
controlled by internal cycling in all three marshes due to prolonged contact times of water with
the subsurface that far exceeds the reaction time of internal cycling pathways.
N Export from Marsh to Open Water
Drainage of N through the creekbank to adjacent open water was a function of the net
advective N fluxes coupled to recharge and interior drainage, combined with internal N cycling
reactions that were uncoupled from advection. TDN-rich porewater was supplied to the
creekbank from the interior of the marsh at low tide, TDN-poor tidal water recharged the marsh
during flooding, and TDN that was internally generated in the subsurface continuously. The
inventory of TDN in the creekbank porewater was large enough that it could not be substantially
changed by the advective fluxes alone. Because the porewater residence times were long relative
to the timescale of N cycling reactions, the porewater N concentration (which was a codeterminant of N flux) was principally modified by inputs from mineralization, removal by plant
uptake and denitrification. Mineralization is the largest component of the internal N cycle,
exhibiting seasonality with increased rates in the spring in some marshes (Tobias et al. 2001a)
and increased rates in the spring and fall in other marshes (Anderson et al. 1997). While we lack
direct measurements of mineralization, its seasonality reported in other marshes is consistent
with quarterly changes in porewater TDN measured in all NRE marshes. Plant uptake also
shows seasonality and variability between dominant vegetation types (Windham-Meyers 2005).
Higher porewater N concentration in the Juncus spp. dominated French Creek and Traps Bay
marshes than in the Spartina spp. dominated Freeman Creek marsh may reflect the slower
growth rate and N uptake of Juncus at those sites. Anthropogenic N inputs do not play a role in
58
porewater N concentration given the low N found in all groundwater samples at all sites.
Denitrification (coupled and direct) is a small component of internal N cycling, especially in
marshes with low ambient nitrate (Anderson et al. 1997, Tobias et al 2001a, Tobias et al. 2003,
Tobias and Neubauer 2009). Based on measured rates (Chapter 2 of this thesis), denitrification
could serve as no more than 5% of a sink for TDN.
The net amount of N internally generated in the creekbank was determined by mass
balancing the internal N sources and sinks with the known advective fluxes of N into and out of
the creekbank (recharge and drainage). In French Creek, contribution from the marsh interior
dominated the creekbank N budget while Traps Bay was more evenly split (55:45) between
contribution from the marsh interior and internal generation. In the up-estuary French Creek site
an average of 537 mg m-2 day-1 was generated within the creekbank while in Traps Bay an
average of 1.6g m-2 day-1 was generated in the creekbank. In the down-estuary Freeman Creek
site, despite shorter porewater residence time, more than 90% of the creekbank N was internally
generated (37.2g m-2 day-1) due largely to the low fluxes of porewater from the interior.
Controls on the Advective Fluxes
The two-order of magnitude range of drainage observed among the marshes more than
spans the drainage rates published for other systems. The 0.5-6 L m shoreline-1 day-1 drainage
rate in Traps Bay was most similar to previously published marsh values (Howes and Goehringer
1994, Harvey and Odum 1990, Tobias et al. 2001b). The French Creek drainage rate was 10%
of those same studies, but Freeman Creek showed almost 10 times the drainage of those
previously documented marshes. One possible explanation of the range of drainage fluxes in the
current study might be differences in hydraulic properties between marshes. However, K values
measured in all the NRE marshes were very similar to those measured by Tobias et al. (2001b)
59
and Harvey and Odum (1990) in two different, and geographically distant, systems. Differences
in tidal amplitude (Fig. 2) between sites may also be a factor. For Harvey and Odum (1990) and
Tobias et al. (2001b), the tidal amplitude was similar to that of Traps Bay, and yielded similar
rates of drainage.
Howes and Goerhinger (1994) studied a marsh with greater amplitude than
Freeman Creek, yet their estimated drainage rates were lower by a factor of 10 compared to what
was calculated for Freeman Creek. The likely explanation for disparity between drainage rates
measured for the NRE and those reported previously might be traced to different approaches in
measuring the drainage flux.
Seepage meters used in some studies (such as Howes and
Goehringer 1994) have been found to underestimate drainage depending upon placement along
the creekbank (Gardner 2005) while salt balance studies (Tobias et al. 2001b) are biased toward
fresh water inputs and drainage is calculated by difference. Few studies (Harvey and Odum
1990, Gardner 2005) have used Darcy-based approaches that are directly comparable to the
present study and those few studies show agreement in drainage estimates for sites of similar
tidal amplitude.
Tidal elevation controls drainage in all marshes. The increase in drainage per unit tidal
elevation below marsh surface is 3 to 4 times the drainage response per unit increase in water
table elevation for Freeman Creek. The increase in drainage per unit tidal elevation below marsh
surface is 50% to 200% of the drainage response per unit increase in water table elevation for
Traps Bay. The increase in drainage per unit tidal elevation below marsh surface is 33% to 66%
of the drainage response per unit increase in water table elevation for French Creek. In contrast
to tidal effects on drainage that can be observed on all time scales (including less than hourly),
upland water table controls on marsh drainage are only observable over a weekly to monthly
scale. Additionally, the French Creek and Traps Bay upland water tables that exert control over
60
drainage were themselves controlled by the tide. Tidal variation accounted for only 2% of
Freeman Creek upland water table elevation but accounted for 15% of French Creek and Traps
Bay upland water table elevation. Particularly for French Creek when examined as a time series,
the upland water table height can be observed tracking high tidal elevations. Therefore, the
periods of high water table elevation that might be equated with a greater slope from the upland
to the marsh were actually coincident with periods of marsh flooding such that only a small
gradient between the upland and marsh existed and upland control on drainage is an artifact of
tidal control on the upland.
Predicting Drainage
Drainage was always correlated to tidal elevation below the marsh platform (Figs. 7-9) at
the creekbank and can serve as a useful predictor of the drainage flux at each site. The slopes of
the regressions below marsh surface at each site represent the response of drainage to lower tidal
elevations (Figs. 11-13). A greater slope indicates a greater response, with Freeman Creek
having the greatest slope and greatest drainage response. Freeman Creek not only had the
greatest response, but also the tightest correlation amongst sites. The high r2 for creekbank
drainage response to tidal elevation indicates that >90% of the variation in drainage is
attributable to tidal elevation changes. The lower r2 in French Creek and Traps Bay is due to the
greater influence of the interior drainage to the creekbank and therefore less than 60% of the
variation in drainage is explained by tidal elevation alone, although the significance of the
correlation is high enough to have confidence in tidal elevation as a proxy of drainage. The
additional variation is explained by water table effects from tidal forcing in French Creek and
combined tidal forcing and precipitation in Traps Bay.
61
While tidal elevation was a good predictor of drainage within a given site, tidal amplitude
may be a better predictor of drainage across the New River Estuary (Fig. 24). Greater tidal
amplitude resulted in greater daily maximum drainage and greater drainage response to
decreased tidal elevation below marsh surface. Since salt marsh elevation is generally optimized
at mean sea level (Morris et al. 2002), larger tidal amplitude affords the opportunity for a lower
low tide end member which translates to greater drainage. The independence of recharge with
depth of tidal flooding indicates that that the amount of tidal amplitude below MSL is the
primary driver of water cycling through the creekbank. A secondary contributor to discharge is
the discharge interface, the cross sectional area through which drainage at the creekbank can
occur measured horizontally as 1m and vertically as the distance between then marsh surface at
the creekbank and the initial slope of the bed of the tidal creek. The area of the discharge
interface increases with greater tidal amplitude. This factor was illustrated among sites with a
discharge area per meter shoreline of 1 m2 at Freeman Creek, 0.33 m2 at Traps Bay, and 0.2 m2
at French Creek. Drainage for the various marshes of the entire New River Estuary could be
estimated using tidal amplitude at the location and tidal elevation below marsh surface and
combined to give an estuary-scale estimate of drainage and its role within the system.
A linear extrapolation between tidal amplitude and drainage (Fig. 24) could be applied to
other systems if more studies at other sites within the tidal amplitude range show a similar
relationship. Given the assumption that tidal amplitude is known and that drainage is tidally
driven in accordance with the measurement presented for the NRE, and with Harvey and Odum
(1990) and Tobias et al. (2001b), a tidal and porewater elevation record of as little as three
months could be used to generate a representative curve of drainage response to tidal elevation,
within a site and of drainage response to tidal amplitude between sites.
62
Marsh Scale N fluxes - Darcy vs Tidal Approaches
The tidal flux experiments yielded estimates of N exchange that were substantially
different in magnitude and in some cases direction relative to the Darcy drainage derived
estimates. Differences between the two methods may be attributable to: 1) Spatial or temporal
variance captured on different scales for water and N; 2) incompletely capturing all routes of N
exchanges; and 3) scaling.
Spatial and Temporal Variance - Darcy
It is possible that the Darcy approach, while providing good temporal coverage, poorly
captured spatial variability that would be important on the integrated marsh scale. Intra-marsh
variability (between transects) in drainage was greatest in Traps Bay because of pronounced
influence of groundwater, which is consistent with Tobias et al. (2001b) but also compromises
the ability to estimate drainage from tidal proxies (Fig. 12)
The Darcy approach afforded the use of auto-logging pressure transducers that provided
good spatial and temporal coverage for Darcy-derived drainage. However, due to the corrosive
nature of the marsh subsurface these instruments operate at the edge of their design and were
subject to failure, which happened with several transducers in this study. Although transducer
failures resulted in data loss and gaps in the record, these failures constituted less than 10% of
the deployment.
Nevertheless, the duration of the record shows a consistent relationship
between drainage and tidal elevation sufficient enough to derive a correlation ranging from 15%
in the smaller amplitude French Creek to 95% in the greater amplitude Freeman Creek. The only
data loss that may have altered the drainage response to tidal elevation would be data lost during
periods of excessively high or excessively low tidal elevation. Considering that drainage was
63
calculated for the entire tidal range, the loss of less than 10% of the data due to instrument
malfunction likely did not affect interpretation of drainage fluxes or their relationship to tidal
dynamics.
Temporal discontinuity in the Darcy N flux arose from pairing a continuous drainage
record with synoptic end-member porewater N concentrations measured seasonally to calculate
the overall N flux. The seasonally measured porewater N-concentrations gave a snapshot of
porewater conditions and assumed there were no daily variations or tide-related variations in N
concentration. Despite the temporal discontinuity between N concentration measurements and
the continuous water flux record, the N concentrations on seasonal scales were within a factor of
2 and monthly N flux varied by a factor of 5-10, reinforcing the conclusion that N flux variance
is primarily driven by water flux variance.
Spatial and Temporal Variance – Tidal Flux
Spatial variation in the tidal water flux was dictated by marsh geomorphology and tidal
characteristics between sites.
Tidal flux studies could only be performed on marsh
geomorphologies that drain to a centralized tidal creek, but the tidal creeks vary considerably in
width, depth, and tidal amplitude
The intent of the tidal flux studies was to provide an alternate estimate of marsh-scale
nitrogen flux to the Darcy-derived estimates that were measured at an expanded marsh scale,
effectively trading an accounting of temporal variability for the appropriate spatial scale. The
temporal variation in porewater drainage (days to weeks) was greater than could be captured by
tidal flux studies, which were point samples in time. Even tidal flux studies that were conducted
on a bi-weekly basis (Wolaver et al. 1998) noted that stochastic events cause large variations in
flux and long term weekly samplings, and Woodwell et al. (1979) showed that despite
64
seasonality there are inter-annual variations. Tidal export on an ebbing tide has been found to be
asymmetrical and variable in relation to tidal amplitude (Boon 1975) which means that the flux
of solutes would also be asymmetrical and variable with tidal amplitude, This tidal asymmetry
leading to a biased solute flux was particularly evident in Freeman Creek (Figs. 28 and 29).
Logistically, there were few opportunities to perform tidal flux studies within the study
period from 2009-2010. A full tidal cycle of 11-13 hours had to be sampled and since these
studies were done on MCBCL the studies had to be standardized for daytime sampling.
Therefore the study could only be done between March 21 and September 21 due to the
photoperiod. When combined with the fact that the sampling had to be scheduled around
MCBCL operations, the number of possible studies was very small. Given the limitations of
tidal flux studies, accessible sites with low variability in tidal amplitude and discharge would be
ideal for these types of studies.
Incomplete Capture of All Routes of N Exchange - Darcy
The Darcy-derived estimates account only for gross subsurface drainage contributions to
the N exchanges. The calculations use porewater N concentrations and measure advective flux,
which does not capture any diffusive exchanges between the marsh and overlying water during
flooded conditions that ultimately drains off the surface and into the open water. This marsh
surface interaction may be a substantial component as evidenced by pulses of high DON in the
May 2010 tidal flux studies that were not seen in the Darcy-based studies. However, Neikirk
(1997) found water-column interactions with the marsh surface also to be an important source of
ammonium to the marsh. Additionally, the Darcy approach can not account for any modification
of N load during discharge at the discharge interface (Gardner 1975).
65
Incomplete Capture of All Routes of N Exchange – Tidal Flux
Water column processes, benthic processes, and runoff from the watershed influenced
nutrient standing stock measurements were an important component of the tidal flux estimate.
Yet, they are unrelated to marsh processes and independent of marsh porewater drainage.
Additionally, the mass flux estimates of import or export are dominated in larger systems like
Freeman Creek by small temporal and spatial variations in concentration applied to coarsely
measured large volumes of tidal creek water inflow or discharge (Boon 1980, Valiela et al.
1980). As described by Wolaver et al. (1988), the water budget of tidal flux studies has the
greatest effect on nutrient transport. Valiela et al. (1980) suggested the best way to alleviate
error was to increase sampling rate, however this study was sampled at the greatest rate possible
and represents the finest temporal resolution of measurements possible.
Scaling - Darcy
For marsh-scale Darcy-derived water flux and nitrogen flux, drainage per meter shoreline
was multiplied by marsh shoreline length. Total shoreline length was an arbitrary determination
that included channels that were 3m in diameter. Estimating the marsh shoreline at a finer
spatial scale for including smaller channels would yield a longer marsh shoreline, and therefore a
larger whole-marsh N-flux. If channels as small as 1m were considered, the marsh shoreline
estimate in Freeman Creek would increase from 9600m estimated with channels >3m to over
15000m and the marsh-integrated N flux would nearly double. Additional consideration must be
given to marsh features such as small peninsulas and isolated islands of marsh that occur within a
marsh system, particularly in the large and complex Freeman Creek, since these features have
only a tidal end-member but are still important because tidally driven recharge and drainage
66
accounts for the majority of the marsh nitrogen flux to surrounding waters. At present, it is
unclear what the appropriate spatial scale for calculating marsh shoreline might be. It is both
unknown and important for appropriate scaling.
Scaling – Tidal Flux
While the tidal flux experiments operate at an integrated marsh scale, estimating the
water flux component was confounded by tidal asymmetry. Additionally, the tidal creeks have
basal flow that exists independent of the tidal forcing and varies with upland water table
elevation and precipitation, which are both variable throughout the year. This factor was
accounted for by standardizing flux calculations based upon salinity (salinity of 0 constituting
watershed input). This was an important factor in Traps Bay with a low volume creek, and less
so in Freeman Creek. Nevertheless, the asymmetry of the flooding versus ebbing water flux
constituted a major difficulty in using tidal flux studies to accurately assess N flux from the
marsh to water at the marsh scale. Combined with the lack of temporal coverage of this
approach, it is difficult to consider this approach as a significant improvement over the Darcyderived estimates. Despite its shortcomings, the Darcy-based approach provides additional
opportunity for predictive applications based on easily measured parameters such as tidal
elevation or tidal amplitude.
67
CHAPTER 2. MARSH-ESTUARY EXCHANGE – ATTENUATION OF DIN FLUXES BY
DENITRIFICATION
INTRODUCTION
It is widely accepted that marshes denitrify (Tobias and Neubauer 2009).
Additional
dissolved inorganic nitrogen (DIN) losses may occur through anaerobic ammonium oxidation
(ANAMMOX), but denitrification dominates (Koop-Jakobsen and Giblin 2009).
Marshes are conducive to denitrification principally because of the high organic carbon
(electron donor) abundance resulting from high rates of primary production. (Dame et al. 2000),
In addition to organic carbon, denitrification requires nitrate and hypoxic (<0.2 mg O2 L-1)
conditions.
Expanded hypoxic zones exist in the marsh subsurface that are suitable for
denitrification. Marsh denitrification rates have been well studied in a variety of settings and
show a large intra- and intermarsh range of rates. (Kaplan et al. 1979, Anderson et al. 1997,
Tobias et al. 2001a, Hammersley and Howes 2005)
Marsh denitrification rates are tightly linked to nitrate availability, which may be derived
directly from the watershed in the form of nitrate-rich groundwater (Kaplan and Valiela 1979,
Giblin and Gaines 1990, Portnoy et al. 1998, Tobias et al. 2001c), delivered by tidal flooding
(particularly in nitrate-rich oligohaline systems (Merrill and Cornwall 2000)). Nitrate is also
supplied by nitrification within the marsh, which tends to be tightly coupled to denitrification
(Hamersley and Howes 2005). The partitioning of the total denitrification rates varying between
coupled
denitrification
(in-situ
nitrification-derived
NO3)
and
direct
denitrification
(allochthonous NO3) appears to be controlled by nitrate concentration (Seitzinger et al. 2006).
Higher NO3 concentrations favor direct denitrification and higher overall total rates (Seitzinger et
al. 2006).
The shallow marsh rhizosphere yields maximum denitrification rates due to close
proximity of abundant denitrifiers to nitrate sources and ample supply of labile organic carbon
69
supplied by roots (Tobias et al. 2001a). With the exception of young marshes, organic matter in
the shallow subsurface exists in sufficient amounts to be non-limiting. However, organic matter
lability may decrease with depth and help to drive the observed exponential decline in potential
denitrification rates with depth below the rhizosphere (Bowden 1986, Tobias et al. 2001a).
Therefore high denitrification rates are encountered in the rhizosphere through which the
porewater transits and drains towards the adjacent estuary. Denitrification has the potential to
attenuate NO3 delivered to the rhizophere from external sources (e.g. groundwater and tidal
infiltration) as well as any of the high NH4 in porewater that is nitrified to NO3 during drainage.
Despite high rates, denitrification may or may not represent a large pathway in the overall
marsh DIN budget (Valiela and Teal 1979, Anderson et al. 1997, Hamersley and Howes 2003).
Assessment of the magnitude of denitrification as a marsh sink for DIN must be placed in the
context of other DIN fluxes, particularly the DIN drainage flux.
The balance between
denitrification and DIN drainage determines a large component of contributing to the overall
effectiveness of the marsh as an attenuator of DIN delivery to the adjacent estuary. Assessing
the rate of denitrification under the current ecosystem state and the total capacity of marsh to
denitrify are both necessary to assess the existing role of denitrification in the marsh DIN source
and sink balance, and the ability of the marsh to buffer the estuary from future increases in
nitrate delivery from the watershed.
Techniques to measure denitrification in general are varied (Groffman et al. 2006) but
many are constrained in the marsh. Membrane inlet mass spec approaches (N2:Ar) are not easily
applied under wetting and drying conditions characteristic of the interidal. Acteylene block
approaches (Seitzinger et al. 2006) exclude coupled denitrification. Marshes with low NO3 can
yield uncharacteristically high rates when supplied with external sources of NO3 required by
70
some in-situ tracer techniques (Addy et al. 2002, Tobias et al. 2001a).
While there is
disagreement on the best denitrification methods (Groffman et al. 2006), use of
15
N tracer in
various forms and applications (Steingruber et al. 2001, Dalsgaard et al. 2005) provides some
flexibility for assessing rates of coupled and direct denitrification under quasi in-situ conditions,
and total marsh capacity for denitrification. The
15
N approach offer high precision with the
ability to isolate individual processes.
In this study we measured rates of N removal via denitrification and ANAMMOX in
three intertidal marshes located along the salinity gradient of the New River Estuary in
southeastern NC. Three methods (isotope pairing technique, potential denitrification, and direct
nitrate addition) were applied to estimate rates of total denitrification, partitioning between
coupled and direct denitrification, ANAMMOX, and denitrification capacity. The derived rates
of overall denitrification-derived sink nature of the marsh was assessed to be combined with
advective marsh N drainage to establish a N budget for the marsh.
METHODS
Coupled denitrification, direct denitrification, denitrification capacity, and ANNAMOX
capacity were assessed in all three marshes seasonally in 2010. Three methods (all using
15
N)
were used to quantify rates and capacities.
Sediment incubations were performed to estimate coupled denitrification using a
15
NH4
spike. Sediment cores (2cm diameter, 3cm deep) were collected at six locations (3 creekside, 3
upland fringe) in each marsh. Cores were collected using cut-off 30mL syringes and transferred
immediately to 40mL amber I-CHEM vials and placed on ice. Four liters of site water was
collected from the adjacent water body at each site and sealed with no headspace.
71
In the lab, three I-CHEM vials from each marsh were loaded with 30mL 2M KCL (prespike). Three I-CHEM vials were spiked with 300µL 10mM
15
NH4 and filled with 30mL 2M
KCl (post-spike). These were placed on a shaker table for 2 hours, filtered, and frozen for DIN
analysis.
Each of the remaining I-CHEM vials was loaded with 300µL 10mM
15
NH4, filled with
corresponding site water, and sealed with no headspace. Prior to sealing, each I-CHEM top was
equipped with a glass bead strung from monofilament that hung halfway down the water column
to serve as a stirring apparatus. The samples were then placed on an orbital shaker table at
60rpm. Samples were killed using 1mL of saturated ZnCl at time intervals of 4, 8, and 18 hours.
At the time of ZnCl addition, sediment was stirred to ensure a complete kill and then recapped
with zero headspace.
Samples were allowed to sit, then 2mL of sample was transferred using a peristaltic pump
to a helium-flushed 12mL exetainer loaded with 50µL KOH. Samples were vortexed and then
run on the isotope ratio mass spectrometer (IRMS) after no less than 30 minutes. The remainder
of the sample was shaken into solution (in the I-CHEM vial), combined with equal parts 2M
KCl, and placed on a shaker table for 2 hours. The samples were then spun down on a centrifuge
at 2000rpm for 10 minutes, filtered, and frozen for DIN analysis.
Isotope analysis was performed using a gas bench on an IRMS to determine N2 volumes
and δ15N. Coupled denitrification rates were calculated according to the following equations:
Coupled denitrification was calculated according the following equations.
The effective enrichment (EE) of the sample was calculated by:
EE = ((CPost-CPre)/CPost)*0.99
(5)
72
where CPost is the NH4 concentration in μM after addition of tracer and CPre is the background
concentration of NH4 in μM before the addition of tracer.
The mole fraction (MF) was calculated by the equation:
MF = ([(δ15N/1000)+1]*0.0036765)/(1+(( δ15N /1000+1)*0.0036765))
(6)
where δ15N is a measured value from the IRMS.
The amount of excess 15N (NE) was calculated by:
NE = (MF - 0.0036630329) * (CN 2* V *2)
(7)
where CN 2 is the concentration of N2 resulting from air equilibration in μM and V is the
overlying water volume in liters.
Denitrification rate (D) was calculated by:
D = [(NE * (1/ EE)) / T ] * (1/ AC)
(8)
where T is the incubation time in hours and AC is the core area (m2).
Additional sediment incubations were performed to estimate coupled and direct
denitrification. Sediment cores were collected as described. In the lab, 3 ICHEM vials from
each marsh were loaded with 225µL 10mM
15
NO3 and filled to the top with corresponding site
water. The overlying water was filtered and frozen for DIN analysis. The remaining ICHEM
vials were loaded with 225µL 10mM 15NO3 and filled to the top with corresponding site water.
Samples were handled, killed, transferred, and extracted for DIN as described above.
Isotope analysis was performed using a gas bench on an IRMS to determine N2 volumes,
29
N and
30
N production, and δ15N.
Denitrification rates were calculated according to the
following equations:
The rate of 29N2 (r29) production was calculated by:
r29 = ((TF 29-T0 29)*(V + (P * S)))/(AC*T)
73
(9)
where TF 29 is the final concentration of 29N2 in μM, T0 29 is the initial concentration of 29N2 in
μM, V is the overlying water volume in liters, P is the porosity of the sediment, S is the sediment
volume in liters, AC is the core area in m2, and T is the incubation time in hours.
The rate of 30N2 (r30) production was calculated by:
r30 = [(TF 30-T0 30)*(V + (P * S))] / (AC*T)
(10)
where TF 30 is the final concentration of 30N2 in μM and T0 30 is the initial concentration of 30N2 in
μM.
The denitrification rate of 15NO3 (D15) was calculated by:
D15 = r29 + 2 * r30
(11)
The denitrification rate of 14NO3 (D14) was calculated by:
D14 = D15 * (r29 / (2 * r30))
(12)
Total denitrification (DTot) was calculated by:
DTot = D14 + D15
(13)
Total water column denitrification (Dcol) was calculated according to:
Dcol = D15 / EE
(14)
where EE is the effective enrichment of the sample.
Coupled denitrification (Dcoupled) was calculated according to:
Dcoupled = (DTot – Dcol) * 2
(15)
Direct denitrification (Ddirect) was calculated according to:
Ddirect = (Dcol * (1 – EE)) * 2
(16)
Denitrification and ANAMMOX capacity were assessed using a sealed tube anaerobic
slurry incubation. The top 3cm of sediment was collected from a creekside location and an
upland fringe location in each marsh. In the lab, each sample was homogenized using a mortar
74
and pestle. Two grams of sediment and a glass bead were added to 12mL exetainers. The
exetainers were capped and incubated overnight in the dark. The exetainers were then flushed
with helium. Two exetainers from each sampling location were frozen to later have porewater
extracted and analyzed for DIN using vanadium reduction.
The exetainers were separated into 4 groups to be loaded with 0.1mL different
concentrations (0.11mM, 0.275mM, 0.825mM, 1.1mM)
15
KNO3 to meet target porewater NO3
concentrations of 10µM, 25µM, 75µM, and 100µM, respectively. All exetainers were then
loaded with 0.1mL 0.55mM NH4Cl and vortexed. The reactions were killed at 0, 1, 2, and 3hour time points using 0.3mL 4M KOH, vortexed, and then ready to run on the IRMS after 2
hours.
Isotope analysis was performed using a gas bench on an IRMS to determine 29N and 30N
production. Denitrification and ANAMMOX rates were calculated according to the following
equation:
Dtotal = P30 * FN-2
(17)
where P30 is 30N2 production and FN is the fraction of 15N in NO3.
ANAMMOX (Atotal) was calculated by:
Atotal = FN-1 * [P29 + 2 * (1- FN-1) * P30]
(18)
where P29 is 29N2 production.
RESULTS
Direct Denitrification
15
N labeled nitrate was added to 2cm thick marsh sediment plugs to assess direct
denitrification. In French Creek, mean creekbank direct denitrification was 1.25 μmoles m-2 hr-1
75
in February and 1.14 μmoles m-2 hr-1 in May. French Creek mean interior direct denitrification
was 1.02 μmoles m-2 hr-1 in February and 1.31 μmoles m-2 hr-1 in May (Fig. 30a). In Traps Bay,
mean creekbank direct denitrification was 1.42 μmoles m-2 hr-1 in February and 2.89 μmoles m-2
hr-1 in May. Traps Bay mean interior direct denitrification was 1.22 μmoles m-2 hr-1 in February
and 1.45 μmoles m-2 hr-1 in May (Fig. 30b).
In Freeman Creek, mean creekbank direct
denitrification was 0.16 μmoles m-2 hr-1 in February and 2.11 μmoles m-2 hr-1 in May. Freeman
Creek mean interior direct denitrification was 0.20 μmoles m-2 hr-1 in February and 0.96 μmoles
m-2 hr-1 in May (Fig. 30c).
Direct denitrification rate was greater in May than in February for all locations at all sites
except French Creek creekbank, which had greater direct denitrification rate in February than
May (Fig. 31a). Direct denitrification was greater in the creekbank than the interior for all
locations in February and May except for French Creek in May (Fig. 31b).
Coupled Denitrification – 15NO3 Addition
Coupled denitrification was calculated from the same incubations as direct denitrification
using isotope pairing technique (IPT) equations presented in Nielsen (1992) and Steingruber et
al. (2001). In French Creek, mean creekbank coupled denitrification was 6.72 μmoles m-2 hr-1 in
February and 4.34 μmoles m-2 hr-1 in May. French Creek mean interior coupled denitrification
was 4.08 μmoles m-2 hr-1 in February and 9.74 μmoles m-2 hr-1 in May (Fig 32b, 32d). In Traps
Bay, mean creekbank coupled denitrification was 14.63 μmoles m-2 hr-1 in February and 23.42
μmoles m-2 hr-1 in May. Traps Bay mean interior coupled denitrification was 11.15 μmoles m-2
hr-1 in February and 4.97 μmoles m-2 hr-1 in May (Fig. 32f, 32h). In Freeman Creek, mean
creekbank coupled denitrification was 3.11 μmoles m-2 hr-1 in February and 8.00 μmoles m-2 hr-1
76
a
Direct denitrification (µmoles N m-2 hr-1)
6
b
6
5
5
4
4
3
3
2
2
1
1
0
0
1
1
c
6
d
6
5
5
4
4
3
3
2
2
1
1
0
0
1
1
e
6
f
6
5
5
4
4
3
3
2
2
1
1
0
0
Creekbank1 Interior
Creekbank1 Interior
February
May
Figure 30. Mean direct denitrification rates from 4 hour incubations for French Creek (a, b),
Traps Bay (c,d), and Freeman Creek (e, f). Black represents creekbank, grey represents interior.
77
May direct denit (µmoles N m-2 hr-1)
4
a
3
2
1
0
Interior direct denit (µmoles N m-2 hr-1)
0
1
2
3
4
-2
February direct denit (µmoles N m hr-1)
4
b
3
2
1
0
0
1
2
3
4
-2
Creekbank direct denit (µmoles N m hr-1)
Figure 31. Comparison of direct denitrification rates of May versus February (a) and marsh
interior versus creekbank (b). Triangles represent French Creek, boxes represent Traps Bay,
diamonds represent Freeman Creek. For May versus February (a), solid symbols represent
creekbank, open symbols represent interior. For interior versus creekbank (b), solid symbols
represent February, open symbols represent May.
78
a
Coupled denitrification (µmoles N m-2 hr-1)
70
b
70
c
70
60
50
60
50
60
50
60
40
30
40
30
40
30
40
20
10
20
10
20
10
20
0
0
0
1
50
30
10
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1
1
e
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f
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60
50
60
40
30
40
20
10
20
g
70
40
30
40
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10
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i
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j
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k
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40
30
40
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40
30
20
10
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February
0
0
1
15NH
4
0
1
1
February
15NO
3
l
70
60
50
0
h
70
60
30
0
1
60
50
50
d
70
May
15NH
4
1
May 15NO3
Figure 32. Mean coupled denitrification rates from 4 hour incubations for French Creek (a-d),
Traps Bay (e-h), and Freeman Creek (i-l). Black represents creekbank sediment incubations,
grey represents interior sediment incubations.
79
in May. Freeman Creek mean interior coupled denitrification was undetectable in February and
3.90 μmoles m-2 hr-1 in May (Fig. 32j, 32l).
Coupled denitrification rate was greater in May than in February for all locations at all
sites (Fig. 33). Coupled denitrification was greater in the creekbank than the interior for all
locations in February and May except for French Creek in May (Fig. 33d).
Coupled Denitrification – 15NH4 Addition
15
N-labeled ammonium was added to separate sediment plugs to provide a second
estimate of coupled nitrification-denitrification exclusive of IPT measurements.
In French
Creek, mean creekbank coupled denitrification was 9.01 μmoles m-2 hr-1 in February and 0.90
μmoles m-2 hr-1 in May. French Creek mean interior coupled denitrification was 9.41 μmoles m-2
hr-1 in February and 5.30 μmoles m-2 hr-1 in May (Fig. 32a, 32c). In Traps Bay, mean creekbank
coupled denitrification was 5.82 μmoles m-2 hr-1 in February and 8.90 μmoles m-2 hr-1 in May.
Traps Bay mean interior coupled denitrification was 6.93 μmoles m-2 hr-1 in February and 3.09
μmoles m-2 hr-1 in May (Fig. 32e, 32g).
In Freeman Creek, mean creekbank coupled
denitrification was 0.64 μmoles m-2 hr-1 in February and 4.95 μmoles m-2 hr-1 in May. Freeman
Creek mean interior coupled denitrification was 0.15 μmoles m-2 hr-1 in February and 1.24
μmoles m-2 hr-1 in May (Fig. 32i, 32k).
The
15
NH4 based coupled denitrification rates were greater in the interior than in the
creekbank for both February and May in French Creek and February in Traps Bay. Coupled
denitrification rates in the creekbank exceeded those in the interior for both February and May in
Freeman Creek and in May at Traps Bay (Fig 33a). Coupled denitrification was greater in May
for all locations at all sites except for the French Creek creekbank, where the rates were
approximately equivalent (Fig. 33c). The 15NH4 approach and IPT-based calculations produced
80
May 15NO3 coupled denit (µmoles N m-2 hr-1)
May 15NH4 coupled denit (µmoles N m-2 hr-1)
a
35
30
30
25
25
20
20
15
15
10
10
5
0
5
Feb
10
15NH
15
4 coupled
20
25
30
denit (µmoles N
m-2
35
Interior 15NO3 coupled denit (µmoles N m-2 hr-1)
0
Interior 15NH4 coupled denit (µmoles N m-2 hr-1)
b
35
hr-1)
c
35
5
0
0
5
10
15
20
25
30
35
30
35
Feb 15NO3 coupled denit (µmoles N m-2 hr-1)
d
30
25
25
20
20
15
15
10
10
5
0
0
5
Creek
10
15NH
4
15
20
25
30
coupled denit (µmoles N
m-2
5
0
35
0
hr-1)
Creek 15NO3 coupled denit (µmoles N m-2 hr-1)
5
10
15
20
25
30
35
Figure 33. Comparison of coupled denitrification rates of marsh May versus February (a,b) and
interior versus creekbank (c,d). Triangles represent French Creek, boxes represent Traps Bay,
diamonds represent Freeman Creek. For May versus February (a,b), solid symbols represent
creekbank, open symbols represent interior. For interior versus creekbank (c,d), solid symbols
represent February, open symbols represent May.
81
similar coupled denitrification rates that, although mean rates differed by a factor of 1.5, were
effectively equivalent within the observed variance.
Comparison of direct denitrification rates versus coupled denitrification rates (15NH4based and IPT) showed that the coupled denitrification rate was greater than direct denitrification
rate for almost all of the sampled locations and times with the exception of the interior of
Freeman Creek in February (Fig. 34).
Coupled denitrification typically exceeded direct
denitrification by a factor of two or greater.
Direct Denitrification Capacity
The strong linear response of direct denitrification rates in sediment slurries to increasing
NO3 concentration indicates high denitrification capacity at the creekbank and marsh interior of
all three sites. The increase in denitrification per μM NO3 increase ranged from 0.02 in the
Freeman Creek creekbank to 0.12 in the Traps Bay creekbank. The increase in denitrification
per μM NO3 increase was more than double in the French Creek creekbank than in the French
Creek interior (Fig. 35a, 35b). The response in the creekbank was 1.25 times that of the response
in the interior in Traps Bay (Fig 35c, 35d). The Freeman Creek creekbank showed the weakest
linear response of all sites, being 25% of the French Creek creekbank response and 20% of the
Traps Bay creekbank response.
The Freeman Creek creekbank denitrification response to
increasing NO3 was 1/5th that of the Freeman Creek interior (Fig. 35e, 35f). For all treatments,
increasing direct denitrification rate occurred up to the highest NO3 concentration (100μM),
indicating that there was no saturation of denitrification kinetics or evidence of co-limiting
substrate. in the 10-100μM range. Anammox rates were low in all locations at all sites at rates of
less than 0.5 nmoles g sed-1 hr-1 measured at higher NO3 concentrations Anammox rates were
82
Creek direct denit (µmoles N m-2 hr-1)
Creek direct denit (µmoles N m-2 hr-1)
a
30
25
20
15
10
5
25
20
15
10
5
0
0
0
5
10
15 20
25
0
30
Creek 15NH4 coupled denit (µmoles N m-2 hr-1)
5
10
15
20
25
30
Creek 15NO3 coupled denit (µmoles N m-2 hr-1)
d
c
30
Interior direct denit (µmoles N m-2 hr-1)
Interior direct denit (µmoles N m-2 hr-1)
b
30
25
20
15
10
5
0
0
5
10
15
20
25
30
30
25
20
15
10
5
0
0
5
10
15
20
25
30
Interior 15NO3 coupled denit (µmoles N m-2 hr-1)
Interior 15NH4 coupled denit (µmoles N m-2 hr-1)
Figure 34. Direct denitrification versus coupled denitrification for all samplings. Triangles
represent French Creek, boxes represent Traps Bay, diamonds represent Freeman Creek. Closed
symbols represent February samples, open symbol represent May samples.
83
y = 0.0728x
0.8
R2 = 0.9887
0.6
0.2
50
100
y = 0.1201x
Traps Bay
14
12
10
8
6
4
2
0
0.4
14
12
10
8
6
4
2
0
14
12
10
8
6
4
2
0
1
2
R = 0.3968
0
150
c
0.2
0
150
1
e
0.8
14
12
10
8
6
4
2
0
0.6
0
y = 0.0157x
0.4
R2 = -0.4888
0.2
50
100
0.4
R2 = 0.9997
0.2
0
14
12
10
8
6
4
2
0
0.4
100
0.6
y = 0.0354x
50
100
150
d
1
0.6
50
1
0.8
0
0.8
0
b
0
150
1
y = 0.0805x
0.6
0.4
0.2
0
0
50
100
y = 0.0924x
R2 = 0.9936
150
1
f
0.8
0.6
0.4
0.2
0
0
Creekbank
0.8
R2 = 0.6676
ANAMMOX (nmoles g sed-1 hr-1)
French Creek
a
0
Freeman Creek
Denitrification rate (nmoles g sed-1 hr-1)
14
12
10
8
6
4
2
0
50
100
150
Interior
NO3 (µM)
Figure 35. Denitrification and ANAMMOX capacity for French Creek creekbank
interior (b), Traps Bay creekbank (c) and interior (d), and Freeman Creek creekbank
interior (f).
Solid symbols denote denitrification, open symbols
ANAMMOX.undetectable at ambient concentrations and ranged from 0-<2%
denitrification at ambient concentrations and at higher NO3 treatments, respectively.
84
(a) and
(e) and
denote
of the
DISCUSSION
Seasonal and Spatial Differences
Total denitrification measured in this study ranged from 1-27 μmoles N m-2 hr-1 and was
overwhelmingly dominated at nearly all times and locations by coupled denitrification. An
increase in total denitrification was seen from February to May in Traps Bay and Freeman Creek
samples while a decrease in total denitrification was seen from February to May for French
Creek samples. This lack of a consistent seasonal pattern between marshes is not surprising
considering variable seasonal results from other studies (Anderson et al. 1997, Eriksson et al.
2003). Kaplan et al. (1979) found that temperature is the primary control on denitrification while
other studies have found that nitrate concentration is the primary control on denitrification
(Eriksson et al. 2003). An increase in ambient nitrate and an increase in temperature were
concurrent in this study and the effects of both could not be isolated. Covariance between
temperature and NO3 as well as other factors that help control coupled denitrification (e.g.
variable mineralization rates in wetlands (Seitzinger 1994) yielded an incomplete picture of
whether denitrification was seasonally different or not in the NRE marshes.
Likewise there was no clear dominance of denitrification in the creekbank versus the
interior. Higher direct denitrification in the creekbank was observed (with one exception; French
Creek in May). Higher rates in the creekbank might be expected where high NO3 in flooding
water is the norm and since a greater volume of tidal water is recharged in the creekbank relative
to the interior. However, high NO3 is not characteristic of the NRE or ICW. It is possible that
higher amounts of plant production in the creekbank might supply a higher concentration of or
more labile DOC. However the results from the denitrification capacity experiments indicate
plenty of useable carbon to drive denitrification at all sites and locations. So it remains unclear
85
why direct denitrification in the creekbank seems to be generally higher relative to the interior.
Regardless, the contribution of direct denitrification relative to coupled denitrification is small
and there was no spatial pattern in coupled denitrification and, by extension, no spatial pattern in
total denitrification. Since coupled denitrification in marshes is limited by the nitrification rate
(Tobias and Neubauer 2009), and nitrification is limited by O2 availability in the subsurface, the
increased recharge of oxygenated tidal water into the creekbank might be expected to enhance
coupled denitrification, or at the least contribute to a larger denitrifying microbial community in
the creekbank. Although the comparison of the creekbank versus interior coupled denitrification
rates observed within a site does not indicates that these O2 delivery mechanisms are contributing
to spatial differences in coupled denitrification rates within a site, coupled denitrification rates at
Traps Bay were 1.5-4 times greater than both French Creek and Freeman Creek (with the
exception of the
15
NH4 experiment in February at French Creek that was higher than both),
indicating that O2 delivery mechanisms differ between sites.
Comparisons to other systems
The range of rates measured in the NRE marshes is comparatively low relative to other
marsh denitrification rates reported (Table 3). Some of the disparity in rate measurements
between the NRE marshes and other systems may be attributable to differences in methods used
to measure denitrification (Groffman et al. 2006).
More likely however, the lower total
denitrification rates in the NRE can be explained primarily by the very small contribution of
direct denitrification that resulted from low ambient NO3 conditions in the NRE. In comparison,
86
Table 3. Denitrification rates reported of other coastal systems
Rate Range (mmol N m-2 hr-1)
System
Denitrification Type
Reference
0.72-1.19
Virginia salt marsh
Total
Anderson et al. 1997
0.00-0 .09
Italian salt marsh
Total
Eriksson et al. 2003
13.00
Freshwater wetlands
Total
Neubauer et al. 2005
0.15
Freshwater wetlands
Total
Hopfensperger et al. 2009
0.10
Global riverine
Total
Seitzinger et al. 2006
0.05
0.17-26.3 × 10-3
Global estuarine
Total
Seitzinger et al. 2006
NC salt marsh
Total
This Study
1.80-17.60
Virginia salt marsh
Direct
Tobias et al. 2001a
0.49-1.50
Virginia salt marsh
Direct
Tobias et al. 2001c
0.20-0.48
-3
0.02-2.89 × 10
New England estuary
Direct
Tobias et al. 2003
NC salt marsh
Direct
This Study
0.017-0.50
New England salt marsh
Coupled
Hamersley and Howes 2005
0.20-0.225
New England salt marsh
Coupled
Hamersley and Howes 2003
0.037 - 0.33
-3
0.15-23.4 × 10
New England salt marsh
Coupled
Tobias et al. 2003
NC salt marsh
Coupled
This study
87
many past studies (Tobias et al 2001a, Tobias et al 2001c, Anderson et al. 1997) were conducted
in systems with high NO3 or under non- NO3 limiting conditions where direct denitrification
rates were high.
The coupled nitrification-denitrification rates measured in this study were also lower
compared to other marsh studies. Coupled denitrification rates derived from both the IPT and
15
NH4-based approaches ranged between 0.2 and 9.4 μmoles N m-2 hr-1, which is 2 to 1000 times
lower compared to New England salt marshes (Table 3). These New England sites are still
supplied with ample watershed NO3 though and may support a larger denitrifying community
that translates into higher coupled denitrification rates as well as higher direct denitrification.
When only low NO3 coastal systems are considered, NRE coupled denitrification rates are still
somewhat low, but are more comparable to other reports (Steingruber et al. 2001, Fennel et al.
2008).
Direct vs. Coupled Denitrification
In the marshes of the NRE, direct denitrification rate was 10-50% that of coupled
denitrification for more than 85% of the measurements and less than 25% for 50% of the
measurements. These ratios are consistent with the observed low NO3 in flooding waters (~1μM
in February, ~3μM in May). Seitzinger et al. (2006) presented a compilation of coupled vs.
direct denitrification ratios across aquatic and marine systems reported as a function of NO3
concentrations.
Coupled- and direct- denitrification were equivalent at ambient NO3
concentration of 10μM. Coupled denitrification was favored at lower NO3 and direct favored at
NO3 concentrations above 10uM. The observed > 80% dominance of coupled, predicted by
Seitzinger et al. (2006) for these NO3 concentrations is consistent with the ratio of coupled
denitrification to direct denitrification of 7:1 measured in the NRE marshes.
88
Low DIN in the NRE (particularly NO3) is the dominant low-flow condition in the NRE
since wastewater treatment improvements were made in the late 1990’s (Mallin et al. 2005).
These conditions were prevalent during the February and May experiments. However pulses of
high NO3 to the NRE accompany high watershed discharge events and a disproportionately
enhanced direct denitrification rate (Fig. 35) would be expected under those conditions.
Denitrification Capacity and Anammox
Exposure of marsh sediment to increased NO3 concentration simulates elevated nitrate loading to
the NRE. The relative response of denitrification is indicative of the marshes’ ability to denitrify
some of those higher NO3 loads considering ambient NO3 concentration in the flooding waters is
currently around 1μM. Increased NO3 concentration resulted in a linear increase in
denitrification rate attributable to enhanced direct denitrification. Rates did not plateau even at
the higher NO3 concentration (100μM) because NO3 and labile organic matter are the limiting
factors regarding denitrification rates. The surface of marshes are organic carbon rich, with
availability well beyond even higher NO3 loads (Tobias et al. 2001a) The relationship between
the increased denitrification rate and the increased concentration shows that given a 100-fold
increase in NO3 load, the marshes would respond with 85-90 fold increase in denitrification for
French Creek, and equivalent 100-fold rise in denitrification for Traps Bay, and an enhanced
125-160 fold increase in denitrification for Freeman Creek. Development of the watershed, an
increased population and wastewater generation, and additional agriculture up-estuary could all
increase NO3 loads, but even at 100-fold increase in current NO3 concentration, it appears that
the marshes of the NRE will accelerate removal of N on an almost 1:1 basis on average. This is
not to say that marshes can remove all of the new N through denitrification, merely that the
proportional amount of removal via denitrification will stay roughly constant relative to the new
89
load.
This response may continue until the large NO3 loads impact the marsh geomorphology
as seen in Drake et al. (2008). Anammox in the marshes of the NRE was negligible, contributing
~1% of the N loss from denitrification. This low contribution of anammox is consistent with the
findings of Koop-Jakobsen and Gibilin (2009), who found anammox to account for less than 3%
of total N2 production. Additionally, the findings from the increasing NO3 incubations using
NRE marsh sediments support the assertions of Koop-Jakobsen and Giblin (2009) that increased
nitrogen loading does not yield a greater proportional contribution of anammox to gaseous N
losses.
90
SUMMARY AND SYNTHESIS
Robust measurements of marsh N drainage and marsh denitrification can be used to
address three broader questions about the role of marshes within the NRE. 1) How significant
are the fluxes of N from the marshes to the NRE / ICW through drainage relative to removal of
N by the marshes through denitrification? 2) How significant are these drainage and
denitrification fluxes relative to other inputs or outputs of the NRE (i.e. river discharge or N
burial through sedimentation); and 3) Is the relative balance of drainage and denitrification likely
to change as marshes respond to climate change?
Scaling Magnitudes of Drainage and Denitrification
To estimate the marsh-scale effects, small scale measurements of drainage (m-1 shoreline)
and denitrification (m-2) of each marsh were scaled to the entire marsh and ultimately to the New
River Estuary system. French Creek, with a marsh shoreline of ~1300 m, accounted for a
drainage driven N loss of 0.05 moles N day-1, or 0.26 kg N yr--1. Traps Bay, with a marsh
shoreline of ~900 m, accounted for a drainage loss of 0.28 moles N day-1, or 1.4 kg N yr--1.
Freeman Creek, with a marsh area of ~9600 m, accounted for a drainage loss of 77.2 moles N
day-1, or 397 kg N yr—1.
French Creek, with a marsh area of ~25,000 m2 and mean interior denitrification rate of
8.31 μmoles m-2 hr-1, accounted for a denitrification driven N loss of 4.99 moles N day-1, or 25.7
kg N yr--1. Traps Bay, with a marsh area of ~14,000 m2 and mean interior denitrification rate of
7.87 μmoles m-2 hr-1, accounted for a denitrification loss of 2.64 moles N day-1, or 13.6 kg N
yr--1. Freeman Creek, with a marsh area of ~450,000 m2 and mean interior denitrification rate of
1.59 μmoles m-2 hr-1, accounted for a denitrification loss of 17.17 moles N day-1, or 88.4 kg N
yr—1. Marsh denitrification removal of N from the NRE system was 1 to 2 orders of magnitude
92
greater than export of N from marshes to adjacent waters via drainage while in the marshes of
the ICW, denitrification removal of N was 20% that of N export via drainage.
Importance of Marsh Drainage and Denitrification Relative to Other Sources and Sinks.
On an annual basis, the N flux of the marshes to the NRE through drainage is low
compared to the input from the mainstem of the NRE and is ~10% of other lateral inputs. The
marsh N flux from the Freeman Creek marsh may be a more important source of N to the ICW,
which has no other significant direct watershed inputs. Much of the N delivery to the NRE is
event-related and increases with pulsed periods of fluvial discharge.
However, porewater
drainage is independent of discharge and contributes a greater percentage of the N standing stock
in the NRE during periods of low discharge. Therefore, porewater drainage may be an important
local source of highly usable reduced N during periods of low fluvial discharge, particularly in
smaller, poorly flushed bodies of water such as Traps Bay and French Creek where adjacent
marshes are draining higher concentrations of N.
Denitrification removes ~700 kg N yr-1 in the NRE and ~350 kg N yr-1 in the ICW, but
is a smaller N sink compared to N burial associated with marsh accretion. Burial can be coarsely
estimated by using measured values of sediment %N (0.3%) and bulk density (2.3 g cm-3) and
assuming that the marshes keep pace with sea level rise at a rate of 2mm per year. Even given
uncertainties in those measurements and assumptions, N burial amounts to 12.6 – 55.2 g N m-2
yr-1 in the NRE marshes and 66.4 g N m-2 yr-1 in the ICW marshes. Extrapolated to the total area
of the NRE and ICW, 8,820 - 38,640 and 112,900 kg N per year are buried in those marshes,
respectively.
The amount of N storage is thus 10 - 55 times larger than removal via
denitrification in the NRE and more than 300 times larger than removal via denitrification in the
ICW. Denitrification is important although it is not the largest flux. However, denitrification
93
represents the only pathway whereby N is truly lost from the system since “buried” N remains
available for remobilization during periods of erosion.
Implications for Marsh N export in a Changing Climate
Marshes are typically situated at sea level and those that can not accrete at a pace greater
than sea level rise will submerge (Morris et al. 2002). The net effect of the processes of marsh
submergence on marsh N export via drainage or marsh removal via denitrification is likely to be
non-linear. As sea level rises and marshes submerge, they will fragment such that shoreline will
increase and marsh area will decrease, thereby increasing the ratio of shoreline length to marsh
area. Since drainage is directly related to shoreline length, the marsh will export more N to the
adjacent waters. As the marsh fragments and submerges, and total denitrification will decrease
with the diminishing marsh area. The marsh will reach an optimum ratio of shoreline to marsh
area (and total shoreline length) at which point drainage is maximized and the drainage to
denitrification ratio is greatest. This change in the source to sink ratio will be attributable solely
to changes in the marsh geomorphology that occurs with fragmentation and alters the shoreline
length to marsh area ratio and will occur even if porewater N concentrations and denitrification
remain constant. Beyond that point of maximized drainage relative to denitrification, the marsh
will submerge until total marsh shoreline also declines with marsh area despite a high
shoreline:marsh area ratio. The loss of total shoreline will result in declining drainage in the
latter stages of submergence, much as it would for loss of denitrification as marsh area drops.
Therefore, the functional relationship between marsh shoreline, marsh area, drainage, and
denitrification must be assessed at each stage of marsh loss in order to quantify the changing
patterns of drainage N export via drainage and denitrification as drowning marshes convert to
open water.
94
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