, Volume 34, Issue 4, pp 641–651

Geochemical, Temperature, and Hydrologic Transport Limitations on Nitrate Retention in Tidal Freshwater Wetlands, Patuxent River, Maryland


    • Department of GeologyUniversity of Maryland
  • Karen Prestegaard
    • Department of GeologyUniversity of Maryland

DOI: 10.1007/s13157-014-0530-5

Cite this article as:
Seldomridge, E. & Prestegaard, K. Wetlands (2014) 34: 641. doi:10.1007/s13157-014-0530-5


Tidal freshwater wetlands receive and retain significant amounts of water, nutrients, and sediment loads from terrestrial watersheds. Wetlands retain nutrients, particularly nitrogen, through microbial processing (e.g. denitrification), plant uptake, and burial. Previous research has provided data on these processes through plot studies and laboratory experiments; however, in situ validation of these results is necessary. Extending the localized measurements to the ecosystem scale requires an understanding of external controls on ecosystem retention processes, such as the determination of whether nitrogen retention is controlled by supply, temperature, or hydrologic transport. These controls were examined through a multi-scale, mass balance approach to measure nitrate retention in tidal freshwater wetlands of the Patuxent River, Maryland. Mass balance measurements of hydrologic and nitrate fluxes were conducted over a 3-year period on a range of marsh sizes. These mass balance results indicate that nitrate retention is not limited by incoming nitrate supply, and is not sensitive to the range of temperatures encountered during the growing season. Nitrate retention data composed of all marsh sites and seasons can be expressed as a simple function of water volume. This result suggests that nitrate retention is principally controlled by hydrologic transport in this tidal freshwater marsh ecosystem.


MarshDenitrificationPatuxent RiverGeomorphologyMass balance


Nitrogen (N) loading from terrestrial landscapes is a major contributor to the development of eutrophic conditions in coastal waters (e.g., Rabalais et al. 2001; Turner and Rabalais 2003). Tidal freshwater wetlands (TFW) function as critical nutrient removal buffers because of their position in the landscape (Simpson et al. 1983). TFW often represent the final opportunity for attenuation of N loads from terrestrial zones before water parcels reach the main estuary (Merrill and Cornwell 2000). Although these wetlands may account for a large proportion of the N removal in coastal nutrient budgets, they occupy a small fraction of the total estuarine area (Simpson et al. 1983; Bowden 1987; Boynton et al. 2008). These ‘hot-spots’ are vegetated, frequently inundated by nitrate-rich water, and have organic carbon supplies necessary for N retention processes. The main N retention processes in the TFW include macrophyte assimilation, microbial denitrification (the reduction of nitrate to N2 or N2O gas), and burial (e.g., Cornwell et al. 1999). The N retention capacity of these sites is influenced by the hydrogeomorphic setting, which affects the amount of N-rich water that is intercepted or received.

The importance of biogeochemical and hydrologic processes on N cycling in TFW is well recognized (e.g., Bowden 1987; Burgin and Hamilton 2007; Ensign et al. 2013). Previous work on controls of N cycling in TFW has focused on identifying and measuring biogeochemical controls (e.g., denitrification rate) through plot or laboratory experiments (e.g., Ensign et al. 2008; Hopfensperger et al. 2009). Although these experiments provide information on in situ processes, results are difficult to extrapolate to field settings due to spatial variations in physiochemical conditions and in the dominant biogeochemical reactions (Schindler 1998; McClain et al. 2003; Davidson and Seitzinger 2006). Spatial variations in physiochemical conditions can be caused by external controls, which are defined as the independent variables that affect N retention processes in a tidal freshwater marsh (e.g., lunar tidal cycle). In the TFWs we identified external controls as: 1) seasonal temperature changes, 2) seasonal and storm-induced variations in stream flow, 3) variations in nitrogen loads from upstream watersheds, and 4) both daily and seasonal variations in tidal stage (Seldomridge 2012). Internal controls arise from interactions among geomorphic, hydrologic, and biological systems within marshes and can be different between marsh systems. The internal controls were: 1) channel inlet and marsh size (investigated in Seldomridge and Prestegaard 2012), and 2) water volume and magnitude of flux as influenced by vegetative roughness and subsequent travel times for a given parcel of water.

The purpose of this study was to examine the influence of environmental variables (external controls) on N removal in TFWs through a multi-scale, mass balance approach. The motivation of this study comes from the Patuxent River Estuary nutrient budget assembled by Boynton et al. (2008). The authors identified the disproportionate importance of TFW of the Patuxent River (one of the major Coastal Plain tributaries to the Chesapeake Bay) as nutrient transformers; however, the budget was assembled by compiling single factor studies (e.g., individual N-loss experiments rather than interactions of N-loss pathways that occur in situ). The approach reported here combines geomorphic, hydrologic, and biogeochemical measurements to evaluate controls on nitrate retention (NR) in TFW of varying size. Mass balance measurements were conducted during spring (high) tides over the growing season from March to October and used to calculate NR rates in marshes from each size class. The data were evaluated to determine whether hydrologic flux, nitrate supply, temperature, pH, and/or dissolved oxygen (D.O.) affected net NR or NR rates. We hypothesized that over the growing season of March to October that: 1) seasonal water temperature changes would affect enzyme kinetics and thus NR rates, 2) that there is a sufficient nitrate supply for NR to occur at its potential rate despite changes in nitrate concentration due to discharge variations or loadings from upstream watersheds, and 3) that pH and dissolved oxygen content of incoming waters would not vary sufficiently to affect nitrate retention within individual marshes.


Study Area

The Patuxent River watershed, located between Washington, D.C. and Baltimore, MD, is 2,260 km2 in total area with 29 km2 of tidal marsh. Historically, the Chesapeake Bay was polluted with nitrogen from distributed sources (fertilizers, urban sources, and septic systems), point sources from sewage and industrial effluents, and atmospheric deposition (Fisher and Oppenheimer 1991; Smullen et al. 1982; Nixon 1987; Fisher et al. 1988; Boynton et al. 1995). A large fraction of the nitrogen load was in the form of inorganic nitrogen from agricultural sources (Boesch et al. 2001). Nitrate loads in the Patuxent River have decreased in the past several decades largely due to reduction in agriculture and control of point sources, but are still significantly higher than watersheds without agricultural or urban land uses (Fisher et al. 2006; Hirsch et al. 2010).

The tidal marshes in this study are contained within Patuxent Wetland Park and Jug Bay Wetlands Sanctuary (Fig. 1). At the time of this study, the TFWs of the Patuxent River (as identified by Boynton et al. 2008) extended 25 river kilometers along the upper Patuxent Estuary (from 39°0′N 76°41′W to 38°43′N 76°41′W). The TFW are composed of hundreds of marshes that range in size from 225 to 675,612 m2 (Seldomridge and Prestegaard 2012). Each tidal freshwater marsh is connected to the tidal Patuxent River through a well-defined inlet channel. Natural levees often border the marsh boundary, particularly in the upper portion of the ecosystem where the study was conducted. These levees prevent direct overbank flooding from the Patuxent River into the adjacent marsh for most tidal stages. Fringing tidal wetlands (those without channel networks) that border the Patuxent River contain a small fraction of the total marsh area. The fringing wetlands are higher in elevation, and subsequently inundated only during the upper 10 % of tidal or flood stages (high tide ± high Patuxent River flow). Water fluxes into tidal marshes were measured at the tidal inlets, which ranged in width from 0.2 to 60 m (Seldomridge and Prestegaard 2012).
Fig. 1

Conceptual diagram of the tidal freshwater marsh ecosystem of the Patuxent River, Chesapeake Bay (upper left). Study sites are located near the head of tide and bracketed by upland and estuarine ecosystems (tidal freshwater wetland area circled). The main inputs of water (Q) and nitrate (N) loads are from the contributing watershed above the head of tide. Hundreds of individual marshes line the Patuxent River, and are responsible for a large proportion of nitrate retention. The geometry of each tidal inlet governs water, sediment, and solute fluxes. Map of Chesapeake Bay courtesy of Jane Thomas, Integration and Application Network, University of Maryland Center for Environmental Science (ian.umces.edu/imagelibrary/)

Site Selection

Geomorphic characteristics of inlet width and marsh surface area were measured for marshes in the TFW ecosystem. Cumulative probability distributions of these geomorphic variables were used to evaluate the size range and frequency of the inlet width and marsh surface areas (Seldomridge and Prestegaard 2012). For this study, three marshes at the upstream end of the TFW ecosystem were chosen. The funnel morphology of the Patuxent River causes this uppermost portion of the TFW to experience the greatest tidal range. Study sites had inlets with widths of 7, 11, and 41 m with associated marsh surface areas of 671, 5,705, and 536,873 m2, respectively. These marsh systems were chosen due to the range of marsh sizes that they represented and the proximity to one another, which made it possible to measure more than one marsh site per tidal cycle during field campaigns.

Tidal Marsh Hydraulic Measurements

Hydraulic variables of width, depth, and velocity were measured to calculate discharge and water volume. For each sampling campaign, tidal stage, maximum velocity (using an electromagnetic current meter), velocity profiles, surface width, and channel cross sections were measured at 10–20 min intervals during tidal cycles with emphasis on the ebb half of each cycle. These data were used to calculate inlet discharge for each time step by establishing a relationship between maximum velocity and average velocity (Chen and Chiu 2002) for each inlet. Discharge (qi) for each time step in the tidal cycle is defined as AiUi, where Ai is the inlet cross sectional area and Ui is the average velocity. Discharge data were integrated over the tidal cycle to determine total water volume (V). A 5 % error in discharge measurement for each time step was propagated through all calculations. This was determined by considering operator error in cross sectional and velocity measurements (Sauer and Meyer 1992), and the error introduced by estimating average velocity from maximum velocity (Chen and Chiu 2002).

Measurements were made over mean high water spring tides (largest tidal range) at the study tidal inlets during the growing season (from March through October) for a 3-year period. The choice of spring tidal conditions was designed to ensure we sampled the range of inundation, D.O., nitrogen concentrations, and temperature.

Marsh and Upper Estuary Water Quality Measurements

Sources of water quality data included publicly accessible U. S. Geological Survey River Input Monitoring Program, samples collected at tidal inlets, and Maryland Department of Natural Resources Eyes on the Bay program. Water quality data in the non-tidal portion of the watershed (upstream of the study sites), were from the U.S. Geological Survey Patuxent River Input Monitoring Program (RIM Station #01594440 near Bowie, MD). The RIM Program reports daily average discharge and N concentration data from bi-monthly samples. These reports provide inflow nitrate concentrations from upstream. Additional discharge values were obtained from the U.S. Geological Survey National Water Information System (USGS NWIS). N concentration data included total Kjeldahl nitrogen and dissolved nitrate (NO2+NO3). For this study, data were acquired for the period from 2008 to 2011.

Field water quality samples were collected from the three study sites (described above). Field samples were collected at tidal inlets during each discharge time step and analyzed for the inorganic nitrogen species NH4-N, NO2-N, and NO3-N (Solorzano 1969; Keefe et al. 2004). Analytical error of ± 2 μM, determined by the mechanical specifications of the equipment, was considered in all NR calculations. Nitrate was the only form of nitrogen that showed significant variation over the tidal cycle, and therefore became the focus of this study (Fig. 2).
Fig. 2

a Conceptual diagram of nitrate dynamics over an ebbing tidal cycle. b Example of water and nitrogen fluxes measured at the tidal inlets for Site 2 on 10/1/2008

Lastly, ambient water quality data from Maryland Department of Natural Resources Eyes on the Bay program were used to examine the influence of environmental variables (temperature, pH, and D.O.) on NR. At the time the study was conducted, Jug Bay Wetlands Sanctuary and Maryland Department of Natural Resources maintained a real-time, permanent (measurements began in 2003), continuous YSI 6600 data logger. The data logger was approximately 3.25 km downstream of the marsh study sites. Measurements of temperature, pH, and D.O. were collected at 15 min intervals and averaged over the incoming tidal cycle (Eyes on the Bay). Maximum ranges are reported for each parameter for the growing season.

Influence of Temperature on NR: Arrhenius Equation

Temperature is an established control on reaction kinetics and it may influence NR during the non-growing season when field mass balance measurements were not made; therefore, water temperatures from the field site were used to evaluate the Arrhenius equation. The Arrhenius equation describes the relationship of the temperature-dependent rate constant, k, of a chemical reaction to temperature T (in Kelvins) and the activation energy, Ea:
$$ k=A{e}^{- Ea/ RT} $$
where A is a coefficient (the pre-factor) and R is the universal gas constant (8.314 × 10−3 kJ mol−1 K−1). A review of the literature provided a range of Ea values (Maag 1997; Holtan-Hartwig et al. 2002; Sheibley et al. 2003; Abdalla et al. 2009); an average Ea value of 47 kJ mol−1 was used. This equation was used to solve for A under a range of documented field conditions to determine the sensitivity of enzyme kinetics to temperature.

Mass Balance Calculations of Nitrate Retention and Nitrate Retention Rates for Individual Marshes

Nitrate retention calculations were made for the 3 study marshes from measurements taken during mean high water springs over the growing season of several years (2008, 2010, and 2011). Nitrate load (NL) was calculated for each time step and summed over the tidal cycle to obtain total nitrate load for flooding or ebbing tides:
$$ NL={\displaystyle \sum}\left({q}_i{C}_i\right) $$

Where NL is nitrate load, qi is discharge for each time step (L s−1), Ci is nitrate concentration for each time step (μM). Net NR was determined by subtracting the outgoing nitrate load from the average incoming nitrate load; incoming nitrate concentrations were measured over half of the incoming tide. The mass balance calculations of NR provided data for individual marshes for a given tidal cycle, time of year, and initial nitrate concentration. No assumption was made of NR processes within each individual marsh, although previous work suggests that marsh surfaces are primary sites for NR (Seldomridge 2009).

In this study, measured nitrate concentrations were relatively constant during the incoming tide (Swarth and Peters 1993; Seldomridge and Prestegaard 2012):
$$ N{L}_{in}= aV $$
Where a is the average incoming nitrate concentration (μM), and V is the total incoming water volume over a tidal cycle (L). Therefore, NR can be expressed as:
$$ NR=a{V}_{in}-b{V}_{out}= cV $$

Where Vin is the incoming water volume (L) and Vout is the outgoing water volume (L), a and b, are incoming and outgoing nitrate concentrations (respectively), and cV is the difference in the flow-weighted nitrate concentrations between incoming and outgoing tides. This equation can be simplified when Vin = Vout, which is the case for tides with minimal water storage and minimal evapotranspirative losses within the marsh. Therefore, field measurements were simplified by measuring only 1–2 h of the incoming tide and the entire ebb tide. The simplifying assumption that total water volume that moved into the tidal channels on a flooding tide was equal to the water volume moving out of the channels on the ebbing tide was made. Previous measurements indicated that these volumes might differ by 2–10 % of the total volume with higher values on the flooding tides during periods of high evapotranspiration, which can be a significant component in the water balance (Hemond et al. 1984). This assumption will therefore tend to underestimate flooding N load and thus N loss. The relationship between NL and NR was used to calculate removal efficiency values expressed as the percentage of the N delivered to these sites (e.g., Nichols 1983; Seitzinger et al. 2006).

The mass balance calculations of NR described above were also used to evaluate NR rates. This was a rate representative of NR processes including denitrification, plant uptake, and burial. Nitrate retention rates were calculated from mass balance studies as:
$$ N{R}_{rate}=\frac{ NR}{T_{inundation}\times {A}_m} $$

Where NRrate is the retention rate (μmol m−2h−1) obtained from mass balance measurements, NR is nitrate retention (moles), Tinundation is the inundation time (hours), which was determined from tidal dynamics and elevation of tidal inlets and marsh surfaces, and Am is the marsh surface area (m2). In this study, Tinundation was estimated from tidal stages measured at tidal inlets and the elevation of marsh surfaces and tidal inlet depths relative to these tidal stages. These mass balance measurements of NR (primarily uptake and denitrification) were then compared with rates determined from laboratory core incubations and field plot studies (Seitzinger 1994; Merrill and Cornwell 2000; Boynton et al. 2008). Error was calculated for each parameter by propagating errors obtained for discharge, geomorphic (see methods of Seldomridge and Prestegaard 2012), and analytical measurements.


Patuxent River Discharge and Nitrate Concentration Data

Patuxent River baseflow was highest during the winter and early spring months when evapotranspiration is at a minimum; baseflow dropped significantly (April through September) during the growing season (Fig. 3c). Peak discharges were associated with storm events, including tropical storms and hurricanes. Annual maximum daily discharge values were 168 m3s−1 in 2008, 145 m3s−1 in 2010, and 354 m3s−1 in 2011 (associated with Hurricane Irene).
Fig. 3

Example of seasonal variations in the Upper Patuxent River in a temperature in 2010, b tidal stage in 2010, c stream flow contributions, and d nitrate concentrations from upstream and field measurements (X). Data sources: temperature and water depth from Eyes on the Bay, Jug Bay permanent monitoring station (a and b); discharge data from USGS NWIS Station 01594440, Patuxent River near Bowie, MD (c); 2008 and 2010 nitrate concentrations from USGS RIM Patuxent River station (d)

Nitrate concentrations in the period 2008 and 2010 showed seasonal variations; high concentrations were associated with late winter-early spring baseflow and lower concentrations occurred during summer months (Fig. 3d). Nitrate was the dominant form of the total N in the upstream measurements (USGS RIM). The nitrate proportion of total N concentration was highest in winter and early spring months (as high as 87.3 % in January 2011) and dropped during summer months as organic nitrogen became a larger proportion of the total N (20 to 40 %).

Relationship of Temperature, D.O., and pH to Nitrate Retention Rate

No significant correlation was found between temperature, D.O., and pH of incoming water parcels and NRrate (Fig. 4; all p-values <0.05). Reported values in Fig. 4 corresponded with average values recorded during field campaigns. During the growing seasons of 2008, 2010, and 2011, temperature ranged from 17.1 °C on 3 April 2010 to 29.2 °C on 18 July 2011. Dissolved oxygen ranged from 5.43 mg L−1 on 14 September 2011 to 11.48 mg L−1 on 18 July 2011. Values of pH ranged from 6.92 S.U. on 14 September 2011 to 7.7 S.U. on 18 July 2011.
Fig. 4

Relationship between upper estuary water quality data (Eyes on the Bay), and NRrate (the rate of collective nitrate retention processes): a temperature, b dissolved oxygen, and c pH for the Patuxent River for 2008–2011 sampling campaigns. Reported water quality data are maximum values measured during the corresponding mass balance measurements. Size of symbol encompasses error values; p-values are approximately >0.95; correlations are not statistically significant

Field water temperatures ranged from 273.15 K (0 °C) in winter to 303.15 K (30 °C) in summer (Fig. 3a; Eyes on the Bay). Temperatures above 10 °C displayed a linear trend with a low slope (from −0.0037 to −0.0088, R2 = 0.96; Fig. 5).
Fig. 5

Range values for A in Arrhenius calculations using an average Ea of 47 kJ mol−1, a range of recorded field water temperatures (0 to 30 °C, and k values dependent on corresponding temperatures (see methods). The coefficient, A, increases the most from 0 to 10 °C; however, above 10 °C the change in values of A is fairly consistant. This suggests the kinetics are not a pronounced limitation during the growing season when temperatures are > 10 °C

Mass Balance Data on Net Nitrate Retention and Retention Rates

Nitrate retention varied with marsh size from less than 1 mol NO3-N per tidal cycle at the smallest site to 2,660 ± 212 mol NO3-N at the largest site (Table 1). To evaluate scaling relationships, these retention values were normalized by total water volume and by marsh surface area for each site. Marsh inundation time was used to convert NR into a retention rate. The smallest site (#1) had the lowest retention values per volume with an average of 6.7 μmol L−1 (NR per volume ranged from 0.6 ± 0.03 to 8.1 ± 0.4 μmol L−1), but highest NRrates with an average of 585.2 μmol m−2 h−1 (NRrate ranged from 28 ± 0.7 to 1,272 ± 69 μmol m−2 h−1). The intermediate site (#2) had intermediate retention values with an average of 12.2 μmol L−1 (NR per volume ranged from 7.1 ± 0.3 to 22.2 ± 1.0 μmol L−1) and rates with an average of 514.2 μmol m−2 h−1 (NRrate ranged 280 ± 15.3 to 822 ± 22.4 μmol m−2 h−1). The large site (#3) had the highest retention values per volume with an average of 19.1 μmol L−1 (NR per volume ranged from 8.8 ± 0.42 to 33.8 ± 1.6 μmol L−1), but the lowest NRrates with an average of 320.6 μmol m−2 h−1 (NRrate ranged from 25 ± 9.3 to 681 ± 29.4 μmol m−2 h−1). On average, the NR per volume values increased with increasing marsh size while average NRrate decreased with increasing marsh size.
Table 1

Field data for Sites 1–3, 2008–2011

Sampling Date

Water Volume (m3)


Total NR (moles)b

\( \raisebox{1ex}{$\boldsymbol{NR}$}\!\left/ \!\raisebox{-1ex}{$\boldsymbol{V}$}\right. \) (μmol L−1)c

NRrate (μmol m−2 h−1)d

Removal efficiency (%)

Site 1


339.6 ± 17

8.1 ± 0.4

2.7 ± 0.34

8.1 ± 0.4

743 ± 40



576.2 ± 28.8

27.4 ± 1.4

4.7 ± 0.004

8.1 ± 0.4

1,272 ± 69



8.2 ± 0.41

0.4 ± 0.02

0.05 ± 0.03

6.1 ± 0.3

28 ± 0.7



235.1 ± 11.8

9.4 ± 0.5

1.1 ± 0.05

4.7 ± 0.2

298 ± 16.2









Site 2


1219.2 ± 61

32.4 ± 1.6

10.6 ± 0.3

8.7 ± 0.4

381 ± 20.7



1459.5 ± 73

84.6 ± 4.2

16.0 ± 7

11.0 ± 0.5

574 ± 31.2



517.6 ± 25.9

30.4 ± 1.5

11.5 ± 2.7

22.2 ± 1.0

822 ± 22.4



1099.5 ± 54.9

44.3 ± 2.2

7.8 ± 5

7.1 ± 0.3

280 ± 15.3









Site 3


16,294 ± 815

544.3 ± 27.2

144.2 ± 6

8.8 ± 0.4

110 ± 22.9



65,708 ± 3,285

5531.9 ± 276.6

1556.2 ± 269.2

23.7 ± 1.1

374 ± 17.6



78,744 ± 3,937

12,477.0 ± 623.8

2659.6 ± 212

33.8 ± 1.6

661 ± 38



72,291.2 ± 3,615

7846.1 ± 392.3

749.0 ± 12

10.4 ± 0.5

186 ± 2.7



38,673.2 ± 1,934

2318.7 ± 115.9

464.6 ± 29.4

12.0 ± 0.6

115 ± 2.5



84,168.0 ± 4,208

6461.6 ± 323.1

1552.4 ± 77.6

18.4 ± 0.9

413 ± 17.5



86,486.9 ± 4,324

5656.2 ± 282.8

2193.6 ± 79.1

25.4 ± 1.2

681 ± 29.4



4,630.3 ± 321

265.1 ± 13.5

92.9 ± 4.6

20.1 ± 0.9

25 ± 9.3









aIncoming nitrate load defined by Eq. 1 (NL = ∑ (qiCi))

bTotal nitrate retention defined by Eq. 2 (NR = NLin − NLout)

cNet nitrate retention per total volume

dNet nitrate retention rate defined by Eq. 3 (\( N{R}_{rate}=\frac{ NR}{T_{inundation}\times {A}_m} \))

eHigh emergent macrophyte growth prevented direct access to channel mouth on 7/18/2011; therefore, discharge and NR values are underestimated

Relationship Between Nitrate Concentration and Nitrate Retention

There was a significant relationship between initial incoming nitrate concentrations and NR (Fig. 6a):
Fig. 6

a Relationship between incoming nitrate concentration (Nin) and nitrate retention (NR) indicates a poor correlation and little increase in NR for nitrate values greater than 60 μM (NR = 3.9Nin4.1, n = 16, R2 = 0.39, p < 0.01). b Relationship between incoming nitrate loads and NR (moles): NR = 0.22NLin, n = 16, R2 = 0.83, p < 0.01. Data points without error bars have error values less than size of symbol

$$ NR=3.9{N_{in}}^{4.1}\left(\mathrm{n}=16,{\mathrm{R}}^2=0.39,\mathrm{p}<0.01\right) $$

Nitrate retention values varied over 3.5 orders of magnitude for very similar values of nitrate in the mid-range of reported values. The general pattern illustrated an increase in NR with nitrate concentration, with little increase in NR when concentrations exceeded 60 μM (Fig. 6a).

Nitrate retention can be expressed as a linear function of incoming nitrate loads (Fig. 6b):
$$ \boldsymbol{NR}=0.22\boldsymbol{N}{\boldsymbol{L}}_{\boldsymbol{in}}\left(\mathrm{n}=16,{\mathrm{R}}^2=0.83,\mathrm{p}<0.01\right) $$

Where NLin is the incoming nitrate load; both parameters are reported in moles. The relationship indicates an overall removal efficiency of 22 % for TFW. This simple linear regression had one data point that fell significantly below the trend. Further examination revealed that this low NR value was associated with low discharge values during a low tidal stage, and should not be compared with the other measurements, which were taken at spring tides. If this outlier is removed, the removal efficiency increases to 24 %. Average removal efficiency values increased with marsh size (Table 1). Average removal efficiency for Site 1 was 18.6 %, Site 2 was 26.7 %, and Site 3 was 25.4 %.

Hydrologic Flux and Nitrate Retention

The volume of water carried into a marsh during a tidal cycle influences total nitrate load and volume appeared to be was a major control on NR. The relationship between tidal water volume and NR for all the marshes defined a single power function relationship (Seldomridge and Prestegaard 2012):
$$ \mathrm{NR}=0.0045{\mathrm{V}}^{1.1}\left(\mathrm{n}=16,{\mathrm{R}}^2=0.98,\mathrm{p}<0.01\right) $$

Where NR is nitrate retention (moles) and V is the water volume (L). This correlation between NR and water volume is evident when NR data were normalized by volume, and although they produced a narrow range of values they indicate that the relationship is not linear (Table 1).

Evaluation of Nitrate Removal Efficiency and Growing Season Variations

Nitrate supply or load into each tidal marsh system was a function of nitrate concentration and discharge integrated over the tidal cycle (Eqs. 2 & 3). The lowest value reported from Site 1 in July was low from dense emergent macrophyte growth in the channel mouth; accessibility to the channel mouth was greatly reduced during this time. The calculated NRrates from the mass balance experiments are shown in Fig. 7. Fewer than half of these NRrates were slightly higher than denitrification rates reported in the literature (Seitzinger 1994; Merrill and Cornwell 2000; Boynton et al. 2008). Rates were highest in May through July, which corresponded with the highest water temperatures, the macrophyte growing season, and likely the highest rates of plant uptake (Fig. 3a & c).
Fig. 7

Seasonal variations in nitrate retention rates calculated from field mass balance data showed slight declines over the growing season. Inset box shows the range of experimental data from literature (Seitzinger 1994; Merrill and Cornwell 2000; Boynton et al. 2008). The high rates during the summer months are likely uptake by plants and/or growth of algal mats. Temperature might limit these values during the winter months (November through March), and transport during the peak summer months. Where error bars are not shown, data point size encompasses error


Influence of Incoming Water Chemistry on NR

There was a positive correlation between incoming nitrate concentration and NR (Fig. 6a, R2 = 0.39, p < 0.01). Comparison with previous research suggests that these measured values of nitrate concentration are higher than required to provide nitrate for microbial processing (e.g., Seitzinger 1988, 1994; Dodds et al. 2002). The shape of the concentration versus NR function (Fig. 6a) suggests that higher concentrations do not increase NR, and may indicate that the system has reached a threshold for nitrate saturation (Ågren and Bosatta 1988; Aber et al. 1989, 1998). In the Patuxent TFW, this threshold appears to be about 60 μM. During this study, the Patuxent TFW received incoming nitrate concentrations higher than 60 μM about 64 % of the time (USGS RIM).

Temperature and D.O. Limitations on Nitrate Retention

Temperature, D.O., and organic carbon availability are known to influence the microbially-mediated transformation of nitrate to gas phases (Patrick and Reddy 1976; Maag 1997; Cornwell et al. 1999). When organic carbon is not limiting, temperature and D.O. are likely controls on NR. Although organic matter content varies with marsh location, in general, organic carbon is not limiting in TFW of the Chesapeake Bay (Pasternack and Brush 1998; Neubauer et al. 2002; Neubauer 2008). In this study, incoming D.O. concentrations showed no correlation with NR; however, these initial oxygen values do not reflect oxygen concentrations at microscale denitrification sites (Burgin and Groffman 2012).

Similarly, temperature is an important control at soil microsites (Burgin and Groffman 2012) because it is an established control on denitrification kinetics. Laboratory experiments have detailed enzyme kinetics of individual loss pathways, such as denitrification (e.g., Betlach and Tiedje 1981); however, once extrinsic effects on denitrification (such as diversity of microbial community, quantity and quality of organic matter, diffusion rates, etc.) are considered, the basic denitrification dynamics are often masked (Seitzinger 1988). The Arrhenius evaluation suggested that NR rates should fall considerably with temperatures less than 5–10 °C (Fig. 5). Temperatures within this range occur for 3–4 months of the year normally from December to March. Above 10 °C, NR retention occurs at an optimum rate (Stanford et al. 1975). In the Patuxent TFW, temperatures are above 10 °C for 7–8 months of the year, which corresponded with the lowest levels of discharge (Fig. 3c). Nitrate retention did not show sensitivity to the narrow range of water temperature variations observed during the growing season, although temperature may slow microbial denitrification during cooler periods not included in this study. During the late fall and winter months when the system receives the highest levels of streamflow, N processing is likely at a seasonal minimum and nitrate may not be processed effectively by the TFW due to temperature constraints and the absence of plant uptake. This unmeasured time period also corresponds with high baseflow discharge and high nitrate concentrations, which suggests that such measurements should be a target for future research.

Hydrologic Transport Limitations on Nitrate Retention

The comparison of nitrate concentrations, loads, and retention suggests NR is limited by the volume of water entering the marsh tidal channel network (Figs. 5, 6 and 7). The differences in the proportions of NR between marshes are principally explained by the differences in magnitude of water fluxes. Some degree of auto-correlation is inherent in this finding (Fig. 6b); however, the relationship between tidal stage and nitrate concentration largely explains the relationship (Fig. 2).

Similarly, Saunders and Kalff (2001) identified this trend across a variety of ecosystems (wetlands, rivers and lakes). In this study, magnitude of water flux was influenced by vegetative roughness from beds of submerged aquatic vegetation, which in turn controls marsh area inundation, travel times and flow paths for a given parcel of water. Although marsh surfaces may have similar denitrification potentials, in situ rates can be limited by local water fluxes (Cooper and Cooke 1984; Hill 1988; Seitzinger et al. 2002; Wollheim et al. 2006). The fine scale marsh hydraulics are examined separately (Seldomridge 2012).

Nitrate Removal Efficiency

Previous studies of nitrate removal efficiency have focused on gauging loss in riparian zones through two methods: (1) measuring nitrate loss in groundwater from upgradient to the stream edge of the buffer (Jordan et al. 1993; Correll 1997; Snyder et al. 1998; Martin et al. 1999), or (2) measuring the travel distance across riparian zones required for nitrate depletion (Lowrance 1998; Wigington et al. 2003). Although these types of studies can provide values of nitrate removal efficiency, they are often not well constrained (Vidon and Dosskey 2008). In this study, overall TFW removal efficiency was 24 % of the incoming nitrate (Eq. 7; Fig 6b), but may increase with higher tides. Tidal prism calculations suggest that in the Patuxent TFW, the same parcel of water may be involved in several tidal cycles (some at night and some during the day) before it exits the freshwater tidal ecosystem (Jenner 2011). This may increase the overall TFW removal efficiency. Removal efficiencies of individual sites ranged from 9.5 to 38.8 % (Table 1), and are within range of previously reported measurements (Nixon and Lee 1986; Cooper 1990; Busnardo et al. 2003). Differences in removal efficiencies may be due to relatively short contact time of water and soil as a result of non-synchronous flooding (Wollheim et al. 2006) and the relatively low proportion of water flooded onto marsh surfaces. These results validated the methodology of direct measurement of water and nutrient fluxes in TFW to calculate removal efficiency.

Similarly, NRrates were similar to, or slightly higher than those reported from other studies including laboratory studies of denitrification using wetland sediment (Fig. 7; Seitzinger 1994; Merrill and Cornwell 2000; Boynton et al. 2008). This comparison suggests that the TFW of the Patuxent River process N near capacity, particularly during fall or spring conditions when macrophyte vegetation does not limit hydrologic fluxes.

Synthesis of Seasonal Controls on Nitrate Retention

In conclusion, growing season measurements indicate the NR in Patuxent TFW is limited by nitrate supply below a threshold level of about 60 μM, above this threshold NR occurs at a maximum potential rate. During the non-storm induced tidal regimes, hydrologic magnitude (tidal stage and resulting water volumes) and transport of available nitrate provide significant controls on NR in the tidal freshwater marshes of the upper Patuxent Estuary; therefore, internal controls (dependent variables) exerted the greatest influence on NR. In the summer, hydrologic transport limitations (tidal + flood stage) are maximized because stream baseflow is low and macrophyte flow resistance is high. High tidal + river stages are most frequent in the late winter and early spring, when temperature is lowest. Thus, cold temperatures may limit NR during winter months when N loads are often at their maximum (Hirsch et al. 2010). The relative importance of temperature and hydrologic controls may be altered by climate change because, among other changes, increasing temperature may extend the period during which denitrification and plant uptake are active (Barendregt and Swarth 2013); however, characterizing these dynamic linkages is best understood through a multi factor study. The Patuxent River TFW are similar to those found along other Coastal Plain Rivers and the study approach could be applied to these other systems.


Dr. Jeffrey Cornwell and Mike Owens, Horn Point Laboratory, University of Maryland Center for Environmental Sciences provided lab assistance and advice that greatly improved this project. We benefited from reviews of an earlier version of this manuscript by Tom Arsuffi, Jeffrey Cornwell, Margaret Palmer, Sujay Kaushal, and Wen-lu Zhu. This research was conducted under an award from the Estuarine Reserves Division, Office of Ocean and Coastal Resource Management, National Ocean Service, National Oceanic and Atmospheric Administration.

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© Society of Wetland Scientists 2014