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Ammonification and denitrification rates in coastal Louisiana bayou sediment and marsh soil: Implications for Mississippi river diversion management
Ecological Engineering 54 (2013) 77–81
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Ammonification and denitrification rates in coastal Louisiana bayou sediment and marsh soil: Implications for Mississippi river diversion management C.M. VanZomeren, J.R. White ∗ , R.D. DeLaune Wetland and Aquatic Biogeochemistry Laboratory, Department of Oceanography and Coastal Sciences, School of the Coast and Environment, Louisiana State University, Baton Rouge, LA 70803, United States
a r t i c l e
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Article history: Received 16 October 2012 Received in revised form 8 January 2013 Accepted 16 January 2013 Available online 24 February 2013 Keywords: Diversion Denitrification Louisiana Mississippi River Nitrate Wetland
a b s t r a c t The Caernarvon Diversion directs Mississippi River water into coastal marshes in the Breton Sound Estuary. Elevated nitrogen levels in the Mississippi River water result in nutrient loading to these coastal marsh systems and consequently to the coastal ocean. The goal of this study is to determine the potential nitrate removal rates for two different substrates. Bayou sediments represent the low flow conditions, when the water is constrained within the canals with high potential for transport to the coastal ocean. The marsh soil represents the high flow diversion events when flood water inundates up into the marshes. We sought to remove the plant effect by using cores containing bayou sediment and marsh soil, removing all roots and flooding with a water column containing 2 mg NO3 -N L−1 . Water column nitrate and ammonium concentration were monitored over 9 d. Net nitrate loss in bayou sediments was 9.5 ± 1.5 mg N m−2 d−1 while the nitrate loss was significantly less at 7.2 ± 0.9 mg N m−2 d−1 in the marsh soil. A comparison of nitrate reduction rates in vegetated and non-vegetated marsh soils indicated that the rate of denitrification increased tenfold in vegetated soils. This increase could be attributed to the “plant effect”. Our results suggest that operating diversions on the high flow end of the spectrum would promote nitrate delivery over the vegetated marsh rather than flowing only through canals. Flooding of the vegetated marsh maximizes the potential for removal of riverine nitrate and limits delivery of nitrate to the coastal ocean, thereby mitigating expressions of eutrophication including algal blooms and hypoxia. © 2013 Elsevier B.V. All rights reserved.
1. Introduction Inorganic nitrogen (N) removal by wetlands is primarily mediated by immobilization and transformation to gaseous forms (N2 , N2 O) through denitrification. Coastal wetlands play important roles in the coastal environment by improving water quality through nutrient abatement and carbon sequestration through organic matter accumulation (DeLaune and White, 2012). Anaerobic conditions and high organic carbon in wetland soils are ideal for nitrate use as an alternate electron donor by denitrifying microbial consortia, thereby removing bioavailable N from the environment (Smith and Tiedje, 1979a; White and Reddy, 1999). As such, assimilation into macrophyte biomass and denitrification are the major nitrate removal pathways in wetland systems. Less significant nitrate removal pathways include phytoplankton uptake, dissimilatory nitrate reduction to NH4 + (DNRA) and anaerobic ammonium oxidation (ANAMMOX) (Reddy and DeLaune, 2008). The significance of
∗ Corresponding author at: 3239 Energy, Coast & Environment, Louisiana State University, United States. Tel.: +1 225 578 8792. E-mail address: [email protected] (J.R. White). 0925-8574/$ – see front matter © 2013 Elsevier B.V. All rights reserved. http://dx.doi.org/10.1016/j.ecoleng.2013.01.029
Mississippi River coastal wetlands serving as nutrient sinks is due to high N from fertilizers linked to intensive agriculture in the Mississippi River basin (Lane et al., 1999). Elevated nutrients, particularly nitrate from agriculture and wastewater discharge (Mitsch et al., 2005; White et al., 2009) in the drainage basin, has resulted in annual algal blooms, including some toxic species, (Bargu et al., 2011) and subsequent hypoxic conditions in the northern Gulf of Mexico (Rabalais et al., 2002). Reduction of nutrient levels, principally nitrate, is necessary for the reduction of eutrophication in the northern Gulf of Mexico (Rabalais et al., 2002). As natural sinks for nutrients, as well as anaerobic conditions and high organic carbon, wetlands have the potential to remove N by denitrification before reaching the Gulf of Mexico, thus helping to alleviate coastal eutrophication (Lane et al., 1999). Restoration projects in the lower Mississippi River include diversions that reintroduce Mississippi River water into adjacent coastal wetlands (Kral et al., 2012). During spring high flow events, the diversions simulate annual spring flooding that occurred prior to installation of levees on the Mississippi River and resultant hydrologic separation of the wetlands from the river. The Caernarvon Diversion is one such diversion and is located downstream of New Orleans, Louisiana and directs river water into the
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C.M. VanZomeren et al. / Ecological Engineering 54 (2013) 77–81 9000 8000
Discharge Rate (ft3s-1)
7000
2008 2009 2010
6000 5000 4000 3000 2000 1000 0
Fig. 1. Caernarvon Mississippi river diversion mean weekly river discharge rates from 2008, 2009, and 2010 (USGS, 2011).
Breton Sound Estuary (Snedden et al., 2007). The wetlands in this estuary are deteriorating due to a number of factors including salt water intrusion, lack of new sediment deposition, land subsidence, and sea level rise (Hyfield et al., 2008). The primary goal of this restoration project is to decrease salt water intrusion as a way to improve fishery and wildlife harvests in the Breton Sound while simultaneously improving deteriorating marshes (Lane et al., 1999). The Caernarvon Diversion has been directing Mississippi River water into the Breton Sound Estuary since 1991, with a maximum discharge of 225 m3 s−1 (8000 cfs). Elevated nitrogen levels in the Mississippi River water result in nutrient loading to this coastal marsh system when the diversion releases water into Breton Sound Estuary (Day et al., 2003). However, discharge in the Breton Sound Estuary varies seasonally and annually based on water levels in the Mississippi River as well as fisheries and wildlife concerns related to water level issues and salinity changes (Fig. 1). Fresh and brackish marshes only receive high nitrate river water when the discharge rate is sufficiently high enough for diverted water in canals (bayous) to flood into marshes. Discharge rates higher than 100 m3 s−1 (3500 cfs) are needed to exceed the capacity for water flow through canals and thus causing sheet flow across the marsh surface (Snedden et al., 2007). With a maximum discharge of 225 m3 s−1 (8000 cfs) into Breton Sound Estuary, at most 44% of all discharged water can flood the marsh (Snedden et al., 2007). When the discharge rate is low, river water remains in two main canals and exogenous nitrate bypasses marshes and is instead directed into the coastal ocean. The Mississippi River has been disconnected from the adjacent riparian wetland areas due to a century long flood protection program. The use of diversions has restored Mississippi River connectivity to adjacent riparian marshes during high spring flood events (Zhang et al., 2012). Coastal wetlands within the Breton Sound Estuary have the potential to mitigate coastal eutrophication by removal of exogenous nitrate, either by plant uptake or denitrification. The goal of this study is to determine the net nitrate removal rates in both bayou sediment and marsh soil. Bayou sediments represent the low flow conditions when the water is constrained within the canals, whereas the marsh soil, with plants and roots removed, represents the high flow diversion events when the water floods the marshes. We hypothesize that a decrease in nitrate concentration will occur faster in the marsh soil than in bayou sediment due to higher carbon content. Furthermore, we compare the denitrification rates of the wetland soil without macrophytes found in this study to a comparable study that evaluated nitrate removal under planted marsh soil conditions.
2. Materials and methods 2.1. Field sampling Bayou sediment and marsh soil were collected from Delacroix (St. Bernard Parish, Louisiana; 29◦ 44.932 N, 89◦ 47.861 W) on January 28, 2011. On the day of sample collection, the mean daily discharge rate of the Caernarvon Diversion was 36 m3 s−1 (1270 cfs). The discharge for January–April 2011 ranged from 13 to 135 m3 s−1 (464–4750 cfs; USGS, 2011). The site is located approximately 16 km from the diversion outfall, with a water depth in the canal approximately 2.5 m and no floodwater apparent in the adjacent marsh. Sediment was collected from the canal bottom center and is classified as bayou sediment in this study. Bayou sediment was collected using a hand dredge to capture the top 20 cm. The marsh soil was collected to include the top 20 cm of the profile and was collected in the marsh adjacent to the bayou station. The marsh site was characterized as an emergent brackish marsh and colonized with monotypic stands of Spartina patens. The sediment and soil were transported to the Wetland and Aquatic Biogeochemistry Laboratory at Louisiana State University and stored refrigerated at 4 ◦ C. Bayou sediment and marsh soil were prepared for use by first removing large organic debris and live roots, respectively. Material was then placed 10.2 cm diameter PVC pipe to a depth of 10 cm for a total of 2 sets of 8 replicate cores (8 bayou sediment and 8 marsh soil). Two groups of 4 cores (8 total replicate cores) were randomly assigned to one of two nitrate treatment concentration groups, either 0.0 (DI water; control) or 2.0 mg NO3 -N L−1 solution (treatment). The 2 mg NO3 -N L−1 concentration was chosen based on observed spring nitrate levels in the Mississippi River (Lane et al., 1999). Nitrate was utilized in this study because >95% of bioavailable N in the Mississippi River is comprised of nitrate (White et al., 2009) with little ammonium present in the river water (Gardner and White, 2010). Each core was flooded with either 0.0 or 2.0 mg NO3 -N L−1 to a depth of 10 cm and changes in nitrate and ammonium concentrations were monitored over 9 d, at 0, 8, 24, 30, 36, 48, 51, 104, 129, 171, 196 h time points. Cores were placed in a 25 ◦ C water bath to maintain a consistent temperature and kept in the dark to prevent the growth of algae. Water column subsamples were taken approximately once daily for the duration of the flood event. Cores were destructively harvested at the end of experiment by removal of the entire sediment or soil core. All samples were stored refrigerated in the dark at 4 ◦ C prior to characterization. Conductivity, redox potential, pH, and dissolved oxygen were monitored at the beginning and again at the end of the experiment time period. Conductivity was monitored in the water column using an Accumet® Basic AB30 Conductivity Meter (Fisher Scientific, Pittsburg, PA) and converted to salinity using the 0.67 conversion factor. Redox potential was measured using a platinum working electrode and saturated calomel (SCE) reference electrode with an Eh conversion factor of +242 (Land et al., 2011). Redox potential was taken at approximately 5 cm soil depth in 8 selected cores, 2 control and 2 treatment cores for each of the bayou sediment and marsh soil (Zhang et al., 2012). The pH in the water column was measured using an Accumet® Research AR25 Dual 133 Channel pH/Ion Meter (Fisher Scientific, Pittsburg, PA). Dissolved oxygen (DO) in the water column was measured using an Accumet® Research AR40 Dissolved Oxygen Meter (Fisher Scientific, Pittsburg, PA). 2.2. Soil/sediment characterization Five subsamples each of the bayou sediment and marsh soil were analyzed for the following physico-chemical characteristics:
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moisture content, total C (TC), total N (TN), KCl extractable NH4 -N, KCl extractable NO3 -N, total P (TP), potentially mineralizable nitrogen (PMN), microbial biomass C (MBC), and microbial biomass N (MBN). Soil moisture was determined by drying a wet subsample at 70 ◦ C to constant weight. Dried soil subsamples were analyzed for TC and TN using an Elemental Combustion System with a detection limit of 0.005 g kg−1 (Costech Analytical Technologies. Inc.,Valencia, CA). Extractable NH4 -N and NO3 -N were measured using 25 mL 2 M KCl soil extractant. Extractable NH4 -N and NO3 -N were analyzed on a SEAL AQ2 Automated Discrete Analyzer (SEAL Analyzer, West Sussex, England; USEPA Methods 132-A Rev. 1 and 103-A Rev. 4, respectively; USEPA, 1993). Method detection limits for NH4 -N was 0.012 mg L−1 and for NO3 -N was 0.014 mg L−1 . The PMN was determined with 25 mL 2 M KCl soil extracts after incubation of 0, 2, 8, and 10 d at 40 ◦ C. The PMN subsamples were subjected to the same EPA methods on the SEAL AQ2 Automated Discrete Analyzer for determination of the extractable NH4 -N. The PMN rate was calculated as the increase in NH4 -N over time by regression (White and Reddy, 2000). Microbial biomass C and N were calculated using the chloroform fumigation-extraction method (Brookes et al., 1985) with adaptations by White and Reddy (2001). Duplicate sets of triplicate 5 g wet weight samples were placed in 25 mL centrifuge tubes. One set was designated as control (non-fumigated) samples and the other set was designated as fumigated samples. Non-fumigated samples were extracted with 25 mL of 0.5 M K2 SO4 , shaken for 30 min then centrifuged. The supernatant was filtered through Whatman #41 filter paper, adjusted to pH < 2, and stored refrigerated in the dark at 4 ◦ C until analysis. Analysis of the supernatant included total organic carbon (TOC) and total dissolved nitrogen (TDN) using a Shimadzu Scientific Instrument TOC-VCSN (Columbia, MD). Fumigated triplicates were placed in a desiccator with chloroform for 24 h. The fumigated set of samples was extracted using the same procedure as the non-fumigated triplicates above. The MBC and MBN were calculated by subtracting the non-fumigated samples from the fumigated samples. Total P was calculated using the TP ashing method after Andersen (1976) for sediment and soil samples. Dried ground sediment was combusted in a muffle furnace (Barnstead Thermolyne 62700 Furnace) at 550 ◦ C for 4 h. Samples were reweighed after burning to determine loss on ignition (LOI). The samples were then digested with 20 mL of 6 M HCl on a hot plate and analyzed for SRP using a SEAL AQ2 Automated Discrete Analyzer (SEAL Analytical, West Sussex, England; Methods 365.1 (USEPA, 1993)). 2.3. Data analysis The effect of nitrate addition between control and treatment cores for the bayou sediment and marsh soil was determined using a Student’s t-test (P < 0.05). Data normality was determined using the Kolmogorov–Smirnov test (˛ = 0.01). Data were logtransformed to fit a normal distribution when necessary. Soil properties compared include bulk density, % moisture, TC, TN, TP, LOI, MBC, MBN, and extractable NH4 + for bayou sediment and marsh soil. Denitrification and N mineralization rate was compared between bayou sediment and marsh soil. 3. Results and discussion 3.1. Soil/sediment characteristics Mean % moisture was 80 ± 0.25% for bayou sediment and 81 ± 0.13% for marsh soil (Table 1). The TC, TN, TP, and LOI were all significantly higher in the marsh soil (165 ± 1.79 g
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Table 1 Soil characteristics for bayou sediment and marsh soil. Data are mean values (n = 2 for bayou PMN and n = 3 for marsh PMN, all other components n = 5) ± sd. Difference letters indicate significant differences between columns at p = 0.5. Soil parameter
Units
Bayou
Marsh
% Moisture TC TN TC:TN TP LOI NH4 -Na MBC MBN PMN∼ Denitrification
% g kg−1 g kg−1
80 ± 0.25a 105 ± 8.69a 7.0 ± 0.66a 15 589 ± 12.2a 22 ± 0.53a 145 ± 8.58a 6.94 ± 0.30a 35 ± 14a 9.63 −9.5 ± 1.5a
81 ± 0.13b 165 ± 1.79b 11 ± 0.10b 14 676 ± 6.85b 32 ± 0.31b 330 ± 7.17b 8.05 ± 0.399b 20 ± 5.8b 9.42 ± 1.62 −7.2 ± 0.9b
a
mg kg−1 % mg kg−1 g kg−1 mg kg−1 mg kg−1 d−1 mg m−2 d−1
Indicates extraction by 2 M KCl.
C kg−1 , 11 ± 0.10 g N kg−1 , 676 ± 6.85 mg P kg−1 , 32 ± 0.31%, respectively) compared to the bayou sediment (105 ± 8.69 g C kg−1 , 7.0 ± 0.66 g N kg−1 and 589 ± 12.2 mg P kg−1 , 22 ± 0.53% respectively). Extractable NH4 -N was approximately 2.3 times higher in the marsh soil than the bayou sediment at 330 ± 7.17 and 145 ± 8.58 mg N kg−1 , respectively. The PMN rate was 9.63 mg N kg−1 d−1 for the bayou sediment and 9.42 ± 1.62 mg N kg−1 d−1 for the marsh soil. In the marsh soil, TC was negatively correlated with % moisture (r = −0.98, n = 5; Table 2). Extractable NH4 + increased as TP increased (r = 0.96) in marsh soils. Redox potential, measured at 5 cm soil depth, was significantly different between bayou sediment and marsh soils at −240 ± 3.4 and −190 ± 31 mV, respectively (p < 0.01). The pH, salinity, and DO, all measured in the water column, were not significantly different between bayou sediment and marsh soil and therefore were averaged together for 7.32 ± 0.11, 0.07 ± 0.09 ppt and 2.97 ± 0.59 mg L−1 , respectively. 3.2. Ammonification and denitrification rates Water column ammonium in the bayou sediment treatment increased at a rate of 0.121 ± 0.008 mg NH4 -N d−1 (Fig. 2a). Water column ammonium increased at a rate of 0.188 ± 0.033 mg NH4 -N d−1 (Fig. 2b) for the marsh soil treatment. The rate of ammonification was significantly lower for the bayou versus marsh (Table 1). On an areal basis, the ammonification rates were 14.8 ± 0.9 and 23.0 ± 4.1 mg NH4 -N m−2 d−1 , respectively for the bayou sediment and marsh soil. Mean nitrate reduction for the bayou sediment treatment was −0.078 ± 0.012 mg NO3 -N d−1 (Fig. 2a). For the marsh soil treatment, nitrate decreased at a rate of −0.059 ± 0.007 mg NO3 -N d−1 . The rate of nitrate removal was significantly greater for bayou sediment versus marsh soil (Table 1; p = 0.04) and was 9.5 ± 1.5 and 7.2 ± 0.9 mg NO3 -N m−2 d−1 , respectively for bayou sediment and marsh soil on an areal basis. We statistically compared the ammonification rates of both control and treatment (added nitrate) cores for both the bayou and marsh treatments in order to determine if the major nitrate reduction pathway was conservative, as in the case for DNRA, or if N was lost from the system by gaseous removal, as for denitrification. If DNRA was a major N transformation mechanism, then the ammonification rates for the nitrate addition treatments would be higher than that of the controls. For the bayou sediment, the rates of ammonium production were not significantly different between control and treatment cores at 14.1 ± 3.3 and 14.8 ± 0.9 mg NH4 N d−1 , respectively. The rates of ammonification were significantly different between control and treatment cores for the marsh soil (34.0 ± 3.6 and 23.0 ± 4.1 mg NH4 -N d−1 ; p = 0.003); however, the
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Table 2 Product-moment correlation coefficients for bayou sediment and marsh soil characteristics. Bold indicates significance at P < 0.05. Bayou sediment TC TN TP LOI MBC MBN NH4 + PMN
% Moisture 0.00 −0.54 −0.62 0.65 −0.36 −0.25 −0.01 0.96
TC
TN
TP
LOI
MBC
MBN
NH4 +
0.79 −0.27 −0.41 0.57 0.13 0.63 0.95
0.01 −0.70 0.83 0.30 0.43 −1.00
−0.65 −0.04 −0.54 −0.49 −0.89
−0.71 0.37 0.22 0.72
0.03 −0.04 −0.69
0.71 −0.04
0.58
Marsh soil
% Moisture
TC
TN
TP
LOI
MBC
MBN
NH4 +
TC TN TP LOI MBC MBN NH4 + PMNa
−0.98 0.23 0.64 −0.64 0.40 −0.13 0.65 0.39
−0.35 −0.48 0.76 −0.43 0.28 −0.53 0.02
−0.06 −0.24 0.54 0.01 0.22 −0.63
0.17 0.43 0.58 0.96 1.00
−0.08 0.80 0.15 0.67
0.03 0.53 0.95
0.63 0.74
0.98
a
n = 3 for PMN, all other components n = 5, r = 0.88.
ammonification rate was higher in the control cores than the treatment cores. Therefore, our data suggest that DNRA was not a major pathway for nitrate reduction in our system. The rates of ammonium production can likely be attributed to N mineralization while the rates of nitrate reduction likely are through denitrification. Soil properties in the bayou sediment and the marsh soil were significantly different because marsh soils were higher in soil organic matter. Extractable NH4 -N was also higher in marsh soil as organic matter serves as the substrate for N mineralization (White and Reddy, 2000). Both bayou sediment and marsh soil treatments experienced a net loss of nitrate over the 9 d study. The rate of nitrate reduction
a
3.0 Nitrate
Ammonium
2.5
mg N L-1
2.0 1.5 1.0 0.5 0.0
b
0
1
2
3
0
1
2
3
4
5
6
7
8
9
3.0 2.5
mg N L-1
2.0 1.5 1.0 0.5 0.0 4
5
6
7
8
9
Time (days) Fig. 2. Mean water column N concentration over 9 d in a bayou sediment (a) and marsh soil (b), presented as mean ± one sd (n = 4).
was significantly higher in the bayou treatment contrary to our original hypothesis. The higher ammonification rates in the marsh treatment may explain the lower rate of nitrate loss in the marsh soil cores, as the mineralized N could have added greater N to the nitrate pool through nitrification (Seitzinger, 1994; Patrick et al., 1996). Significantly higher TC and extractable NH4 -N in marsh soils support N mineralization in the marsh soils. Increasing water column ammonium concentrations also support higher N mineralization in marsh soil (23.0 ± 4.1 mg NH4 -N m−2 d−1 ) than the bayou sediment (14.8 ± 0.9 mg NH4 -N m−2 d−1 ). Adjacent zones of aerobic and anaerobic conditions are needed for coupling of nitrification and denitrification to occur (White and Reddy, 2003). Nitrification of ammonium to nitrate occurs only under aerobic conditions. Water column DO levels of 2.97 ± 0.59 mg L−1 suggest conditions for nitrification to occur were present. Conversely, denitrification only occurs under anaerobic conditions. Redox measurements of −240 ± 3.4 and −190 ± 31 mV at 5 cm below the sediment/soil surface suggest conditions existed for denitrification to occur (Patrick et al., 1996). Diffusion of ammonium from N mineralization in sediment/soil into the oxygenated water column increases nitrate concentration in the water column (Reddy et al., 1978). In addition, bayou sediments may have had a higher initial active microbial consortium as the river water with associated nitrate was present only in the bayous while the marsh was not flooded by river water at the time of collection. Research has shown that wetland soils/sediments previously in contact with nitrate have a higher active denitrifying enzyme activity (White and Reddy, 1999) and that in general, the higher the concentration of nitrate, the higher the initial enzyme activity (Gardner and White, 2010). The marsh denitrification rates are low, a possible artifact of root removal. Consequently, we compared the rate of nitrate loss for the marsh soil (this study) to a published companion study where the rate of nitrate loss was monitored in a marsh soil colonized by macrophytes at the same site. Marsh soil planted with macrophytes reduced a 2 mg NO3 -N L−1 water column to below detection within 12 h, as opposed to this study, where the marsh soil without plants only reduced about half of the 2 mg NO3 -N L−1 water column over 9 d (VanZomeren et al., 2012). The previously published study partitioned the 15 N-NO3 into soil, plant and gaseous loss components. The rates of nitrate loss attributed to denitrification in that study was over an order of magnitude greater than the denitrification rates seen in this study.
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This comparison clearly demonstrates macrophytes critical role in promoting denitrification (Smith and Tiedje, 1979b). In addition, there was no ammonium release from the macrophyte treatment compared to significant ammonium released in our study with no plants. Consequently, the presence of macrophytes not only increased denitrification but also prevented the release of ammonium due to immobilization. The management implications for this study relate to the operation of the large-scale Mississippi River diversions. Although Mississippi River diversions were built to reduce salinity and enhance wildlife and fisheries in receiving wetlands, it has become clear that these diversions can also help mitigate nutrient loads, particularly nitrate, in diverted Mississippi River water before discharge into the Gulf of Mexico. This, however, can only be achieved if discharge rates are managed to enhance denitrification potential in the receiving basin. At low flow, the high nitrate river water passes through the bayou canals with low denitrification potential and therefore delivers nitrate out to the coastal ocean. At higher flow, the river water floods the vegetated marsh areas which increases the removal rate of nitrate and reduces nitrate export to the coastal ocean. Concerns over possible reductions in marsh stability due to soil C oxidation through denitrification were not supported by stoichiometric calculations in the companion study (VanZomeren et al., 2012), further supporting the use of higher diversion flows to enhance denitrification potential and removal of nitrate before discharging to the coastal ocean. 4. Conclusion Net nitrate removal was significantly higher in bayou sediments compared with marsh soil when roots are removed (9.5 ± 1.5 and 7.2 ± 0.9 mg NO3 -N m−2 d−1 , respectively). The ammonification rate is significantly lower in bayou sediments than marsh soil (14.8 ± 0.9 and 23.0 ± 4.1 mg NH4 -N m−2 d−1 , respectively). This ammonium production suggests that both bayou sediments and marsh soil can potentially be a source of nitrate through coupled N mineralization and nitrification, thereby decreasing potential nitrate removal capacity by denitrification. Comparison of our rates with a companion study that included plants suggest nitrate reduction rates in vegetated and non-vegetated marsh soils indicated that the rate of denitrification increased tenfold in vegetated soils as well as assimilation of released ammonium into biomass. Management of diversion flow rate is critical for enhancing the potential reduction of nitrate by soil denitrification. Lower flows do not allow the vegetated marsh to be flooded, resulting in a lower denitrification potential and less nitrate removal. Therefore, the operation of river diversions should be maintained on the high flow end of the spectrum to assure that nitrate is delivered into the vegetated marsh, if the goal is maximum nitrate removal from river water. Acknowledgments We would like to acknowledge Louisiana Sea Grant Program and the Louisiana Office of Coastal Protection and Restoration for providing funding for this research by way of a Coastal Sciences Assistantship. We also would like to thank Nathan Nguyen and Anthony Nguyen for laboratory assistance.
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Methods for the Determination of Inorganic Substances in Environmental Samples, EPA/600/R-93/100, Washington, D.C. USGS, 2011. http://waterdata.usgs.gov/usa/nwis/uv?site no=295124089542100 VanZomeren, C.M., White, J.R., DeLaune, R.D., 2012. Fate of nitrate in vegetated brackish coastal marsh. Soil Sci. Soc. Am. J. 76, 1919–1927. White, J.R., Fulweiler, R.W., Li, C.Y., Bargu, S., Walker, N.D., Twilley, R.R., Green, S.E., 2009. Mississippi river flood of 2008: observations of a large freshwater diversion on physical, chemical, and biological characterisitics of a shallow estuarine lake. Environ. Sci. Technol. 43 (15), 5599–5604. White, J.R., Reddy, K.R., 1999. Influence of nitrate and phosphorus loading on denitrifying enzyme activity in Everglades wetland soils. Soil Sci. Soc. Am. J. 63, 1945–1954. White, J.R., Reddy, K.R., 2000. Influence of phosphorus loading on organic nitrogen mineralization of Everglades soils. Soil Sci. Soc. Am. J. 64, 1525–1534. White, J.R., Reddy, K.R., 2001. Influence of selected inorganic electron acceptors on organic nitrogen mineralization in Everglades soils. Soil Sci. Soc. Am. J. 65, 941–948. White, J.R., Reddy, K.R., 2003. Nitrification and denitrification rates of Everglades wetland soils along a phosphorus-impacted gradient. J. Environ. Qual. 32, 2436–2443. Zhang, W., White, J.R., DeLaune, R.D., 2012. Diverted Mississippi river sediment as a potential phosphorus source affecting coastal Louisiana water quality. J. Fresh. Ecol. 27 (4), 575–586.
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Ammonification and denitrification rates in coastal Louisiana bayou sediment and marsh soil: Implications for Mississippi river diversion management C.M. VanZomeren, J.R. White ∗ , R.D. DeLaune Wetland and Aquatic Biogeochemistry Laboratory, Department of Oceanography and Coastal Sciences, School of the Coast and Environment, Louisiana State University, Baton Rouge, LA 70803, United States
a r t i c l e
i n f o
Article history: Received 16 October 2012 Received in revised form 8 January 2013 Accepted 16 January 2013 Available online 24 February 2013 Keywords: Diversion Denitrification Louisiana Mississippi River Nitrate Wetland
a b s t r a c t The Caernarvon Diversion directs Mississippi River water into coastal marshes in the Breton Sound Estuary. Elevated nitrogen levels in the Mississippi River water result in nutrient loading to these coastal marsh systems and consequently to the coastal ocean. The goal of this study is to determine the potential nitrate removal rates for two different substrates. Bayou sediments represent the low flow conditions, when the water is constrained within the canals with high potential for transport to the coastal ocean. The marsh soil represents the high flow diversion events when flood water inundates up into the marshes. We sought to remove the plant effect by using cores containing bayou sediment and marsh soil, removing all roots and flooding with a water column containing 2 mg NO3 -N L−1 . Water column nitrate and ammonium concentration were monitored over 9 d. Net nitrate loss in bayou sediments was 9.5 ± 1.5 mg N m−2 d−1 while the nitrate loss was significantly less at 7.2 ± 0.9 mg N m−2 d−1 in the marsh soil. A comparison of nitrate reduction rates in vegetated and non-vegetated marsh soils indicated that the rate of denitrification increased tenfold in vegetated soils. This increase could be attributed to the “plant effect”. Our results suggest that operating diversions on the high flow end of the spectrum would promote nitrate delivery over the vegetated marsh rather than flowing only through canals. Flooding of the vegetated marsh maximizes the potential for removal of riverine nitrate and limits delivery of nitrate to the coastal ocean, thereby mitigating expressions of eutrophication including algal blooms and hypoxia. © 2013 Elsevier B.V. All rights reserved.
1. Introduction Inorganic nitrogen (N) removal by wetlands is primarily mediated by immobilization and transformation to gaseous forms (N2 , N2 O) through denitrification. Coastal wetlands play important roles in the coastal environment by improving water quality through nutrient abatement and carbon sequestration through organic matter accumulation (DeLaune and White, 2012). Anaerobic conditions and high organic carbon in wetland soils are ideal for nitrate use as an alternate electron donor by denitrifying microbial consortia, thereby removing bioavailable N from the environment (Smith and Tiedje, 1979a; White and Reddy, 1999). As such, assimilation into macrophyte biomass and denitrification are the major nitrate removal pathways in wetland systems. Less significant nitrate removal pathways include phytoplankton uptake, dissimilatory nitrate reduction to NH4 + (DNRA) and anaerobic ammonium oxidation (ANAMMOX) (Reddy and DeLaune, 2008). The significance of
∗ Corresponding author at: 3239 Energy, Coast & Environment, Louisiana State University, United States. Tel.: +1 225 578 8792. E-mail address: [email protected] (J.R. White). 0925-8574/$ – see front matter © 2013 Elsevier B.V. All rights reserved. http://dx.doi.org/10.1016/j.ecoleng.2013.01.029
Mississippi River coastal wetlands serving as nutrient sinks is due to high N from fertilizers linked to intensive agriculture in the Mississippi River basin (Lane et al., 1999). Elevated nutrients, particularly nitrate from agriculture and wastewater discharge (Mitsch et al., 2005; White et al., 2009) in the drainage basin, has resulted in annual algal blooms, including some toxic species, (Bargu et al., 2011) and subsequent hypoxic conditions in the northern Gulf of Mexico (Rabalais et al., 2002). Reduction of nutrient levels, principally nitrate, is necessary for the reduction of eutrophication in the northern Gulf of Mexico (Rabalais et al., 2002). As natural sinks for nutrients, as well as anaerobic conditions and high organic carbon, wetlands have the potential to remove N by denitrification before reaching the Gulf of Mexico, thus helping to alleviate coastal eutrophication (Lane et al., 1999). Restoration projects in the lower Mississippi River include diversions that reintroduce Mississippi River water into adjacent coastal wetlands (Kral et al., 2012). During spring high flow events, the diversions simulate annual spring flooding that occurred prior to installation of levees on the Mississippi River and resultant hydrologic separation of the wetlands from the river. The Caernarvon Diversion is one such diversion and is located downstream of New Orleans, Louisiana and directs river water into the
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C.M. VanZomeren et al. / Ecological Engineering 54 (2013) 77–81 9000 8000
Discharge Rate (ft3s-1)
7000
2008 2009 2010
6000 5000 4000 3000 2000 1000 0
Fig. 1. Caernarvon Mississippi river diversion mean weekly river discharge rates from 2008, 2009, and 2010 (USGS, 2011).
Breton Sound Estuary (Snedden et al., 2007). The wetlands in this estuary are deteriorating due to a number of factors including salt water intrusion, lack of new sediment deposition, land subsidence, and sea level rise (Hyfield et al., 2008). The primary goal of this restoration project is to decrease salt water intrusion as a way to improve fishery and wildlife harvests in the Breton Sound while simultaneously improving deteriorating marshes (Lane et al., 1999). The Caernarvon Diversion has been directing Mississippi River water into the Breton Sound Estuary since 1991, with a maximum discharge of 225 m3 s−1 (8000 cfs). Elevated nitrogen levels in the Mississippi River water result in nutrient loading to this coastal marsh system when the diversion releases water into Breton Sound Estuary (Day et al., 2003). However, discharge in the Breton Sound Estuary varies seasonally and annually based on water levels in the Mississippi River as well as fisheries and wildlife concerns related to water level issues and salinity changes (Fig. 1). Fresh and brackish marshes only receive high nitrate river water when the discharge rate is sufficiently high enough for diverted water in canals (bayous) to flood into marshes. Discharge rates higher than 100 m3 s−1 (3500 cfs) are needed to exceed the capacity for water flow through canals and thus causing sheet flow across the marsh surface (Snedden et al., 2007). With a maximum discharge of 225 m3 s−1 (8000 cfs) into Breton Sound Estuary, at most 44% of all discharged water can flood the marsh (Snedden et al., 2007). When the discharge rate is low, river water remains in two main canals and exogenous nitrate bypasses marshes and is instead directed into the coastal ocean. The Mississippi River has been disconnected from the adjacent riparian wetland areas due to a century long flood protection program. The use of diversions has restored Mississippi River connectivity to adjacent riparian marshes during high spring flood events (Zhang et al., 2012). Coastal wetlands within the Breton Sound Estuary have the potential to mitigate coastal eutrophication by removal of exogenous nitrate, either by plant uptake or denitrification. The goal of this study is to determine the net nitrate removal rates in both bayou sediment and marsh soil. Bayou sediments represent the low flow conditions when the water is constrained within the canals, whereas the marsh soil, with plants and roots removed, represents the high flow diversion events when the water floods the marshes. We hypothesize that a decrease in nitrate concentration will occur faster in the marsh soil than in bayou sediment due to higher carbon content. Furthermore, we compare the denitrification rates of the wetland soil without macrophytes found in this study to a comparable study that evaluated nitrate removal under planted marsh soil conditions.
2. Materials and methods 2.1. Field sampling Bayou sediment and marsh soil were collected from Delacroix (St. Bernard Parish, Louisiana; 29◦ 44.932 N, 89◦ 47.861 W) on January 28, 2011. On the day of sample collection, the mean daily discharge rate of the Caernarvon Diversion was 36 m3 s−1 (1270 cfs). The discharge for January–April 2011 ranged from 13 to 135 m3 s−1 (464–4750 cfs; USGS, 2011). The site is located approximately 16 km from the diversion outfall, with a water depth in the canal approximately 2.5 m and no floodwater apparent in the adjacent marsh. Sediment was collected from the canal bottom center and is classified as bayou sediment in this study. Bayou sediment was collected using a hand dredge to capture the top 20 cm. The marsh soil was collected to include the top 20 cm of the profile and was collected in the marsh adjacent to the bayou station. The marsh site was characterized as an emergent brackish marsh and colonized with monotypic stands of Spartina patens. The sediment and soil were transported to the Wetland and Aquatic Biogeochemistry Laboratory at Louisiana State University and stored refrigerated at 4 ◦ C. Bayou sediment and marsh soil were prepared for use by first removing large organic debris and live roots, respectively. Material was then placed 10.2 cm diameter PVC pipe to a depth of 10 cm for a total of 2 sets of 8 replicate cores (8 bayou sediment and 8 marsh soil). Two groups of 4 cores (8 total replicate cores) were randomly assigned to one of two nitrate treatment concentration groups, either 0.0 (DI water; control) or 2.0 mg NO3 -N L−1 solution (treatment). The 2 mg NO3 -N L−1 concentration was chosen based on observed spring nitrate levels in the Mississippi River (Lane et al., 1999). Nitrate was utilized in this study because >95% of bioavailable N in the Mississippi River is comprised of nitrate (White et al., 2009) with little ammonium present in the river water (Gardner and White, 2010). Each core was flooded with either 0.0 or 2.0 mg NO3 -N L−1 to a depth of 10 cm and changes in nitrate and ammonium concentrations were monitored over 9 d, at 0, 8, 24, 30, 36, 48, 51, 104, 129, 171, 196 h time points. Cores were placed in a 25 ◦ C water bath to maintain a consistent temperature and kept in the dark to prevent the growth of algae. Water column subsamples were taken approximately once daily for the duration of the flood event. Cores were destructively harvested at the end of experiment by removal of the entire sediment or soil core. All samples were stored refrigerated in the dark at 4 ◦ C prior to characterization. Conductivity, redox potential, pH, and dissolved oxygen were monitored at the beginning and again at the end of the experiment time period. Conductivity was monitored in the water column using an Accumet® Basic AB30 Conductivity Meter (Fisher Scientific, Pittsburg, PA) and converted to salinity using the 0.67 conversion factor. Redox potential was measured using a platinum working electrode and saturated calomel (SCE) reference electrode with an Eh conversion factor of +242 (Land et al., 2011). Redox potential was taken at approximately 5 cm soil depth in 8 selected cores, 2 control and 2 treatment cores for each of the bayou sediment and marsh soil (Zhang et al., 2012). The pH in the water column was measured using an Accumet® Research AR25 Dual 133 Channel pH/Ion Meter (Fisher Scientific, Pittsburg, PA). Dissolved oxygen (DO) in the water column was measured using an Accumet® Research AR40 Dissolved Oxygen Meter (Fisher Scientific, Pittsburg, PA). 2.2. Soil/sediment characterization Five subsamples each of the bayou sediment and marsh soil were analyzed for the following physico-chemical characteristics:
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moisture content, total C (TC), total N (TN), KCl extractable NH4 -N, KCl extractable NO3 -N, total P (TP), potentially mineralizable nitrogen (PMN), microbial biomass C (MBC), and microbial biomass N (MBN). Soil moisture was determined by drying a wet subsample at 70 ◦ C to constant weight. Dried soil subsamples were analyzed for TC and TN using an Elemental Combustion System with a detection limit of 0.005 g kg−1 (Costech Analytical Technologies. Inc.,Valencia, CA). Extractable NH4 -N and NO3 -N were measured using 25 mL 2 M KCl soil extractant. Extractable NH4 -N and NO3 -N were analyzed on a SEAL AQ2 Automated Discrete Analyzer (SEAL Analyzer, West Sussex, England; USEPA Methods 132-A Rev. 1 and 103-A Rev. 4, respectively; USEPA, 1993). Method detection limits for NH4 -N was 0.012 mg L−1 and for NO3 -N was 0.014 mg L−1 . The PMN was determined with 25 mL 2 M KCl soil extracts after incubation of 0, 2, 8, and 10 d at 40 ◦ C. The PMN subsamples were subjected to the same EPA methods on the SEAL AQ2 Automated Discrete Analyzer for determination of the extractable NH4 -N. The PMN rate was calculated as the increase in NH4 -N over time by regression (White and Reddy, 2000). Microbial biomass C and N were calculated using the chloroform fumigation-extraction method (Brookes et al., 1985) with adaptations by White and Reddy (2001). Duplicate sets of triplicate 5 g wet weight samples were placed in 25 mL centrifuge tubes. One set was designated as control (non-fumigated) samples and the other set was designated as fumigated samples. Non-fumigated samples were extracted with 25 mL of 0.5 M K2 SO4 , shaken for 30 min then centrifuged. The supernatant was filtered through Whatman #41 filter paper, adjusted to pH < 2, and stored refrigerated in the dark at 4 ◦ C until analysis. Analysis of the supernatant included total organic carbon (TOC) and total dissolved nitrogen (TDN) using a Shimadzu Scientific Instrument TOC-VCSN (Columbia, MD). Fumigated triplicates were placed in a desiccator with chloroform for 24 h. The fumigated set of samples was extracted using the same procedure as the non-fumigated triplicates above. The MBC and MBN were calculated by subtracting the non-fumigated samples from the fumigated samples. Total P was calculated using the TP ashing method after Andersen (1976) for sediment and soil samples. Dried ground sediment was combusted in a muffle furnace (Barnstead Thermolyne 62700 Furnace) at 550 ◦ C for 4 h. Samples were reweighed after burning to determine loss on ignition (LOI). The samples were then digested with 20 mL of 6 M HCl on a hot plate and analyzed for SRP using a SEAL AQ2 Automated Discrete Analyzer (SEAL Analytical, West Sussex, England; Methods 365.1 (USEPA, 1993)). 2.3. Data analysis The effect of nitrate addition between control and treatment cores for the bayou sediment and marsh soil was determined using a Student’s t-test (P < 0.05). Data normality was determined using the Kolmogorov–Smirnov test (˛ = 0.01). Data were logtransformed to fit a normal distribution when necessary. Soil properties compared include bulk density, % moisture, TC, TN, TP, LOI, MBC, MBN, and extractable NH4 + for bayou sediment and marsh soil. Denitrification and N mineralization rate was compared between bayou sediment and marsh soil. 3. Results and discussion 3.1. Soil/sediment characteristics Mean % moisture was 80 ± 0.25% for bayou sediment and 81 ± 0.13% for marsh soil (Table 1). The TC, TN, TP, and LOI were all significantly higher in the marsh soil (165 ± 1.79 g
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Table 1 Soil characteristics for bayou sediment and marsh soil. Data are mean values (n = 2 for bayou PMN and n = 3 for marsh PMN, all other components n = 5) ± sd. Difference letters indicate significant differences between columns at p = 0.5. Soil parameter
Units
Bayou
Marsh
% Moisture TC TN TC:TN TP LOI NH4 -Na MBC MBN PMN∼ Denitrification
% g kg−1 g kg−1
80 ± 0.25a 105 ± 8.69a 7.0 ± 0.66a 15 589 ± 12.2a 22 ± 0.53a 145 ± 8.58a 6.94 ± 0.30a 35 ± 14a 9.63 −9.5 ± 1.5a
81 ± 0.13b 165 ± 1.79b 11 ± 0.10b 14 676 ± 6.85b 32 ± 0.31b 330 ± 7.17b 8.05 ± 0.399b 20 ± 5.8b 9.42 ± 1.62 −7.2 ± 0.9b
a
mg kg−1 % mg kg−1 g kg−1 mg kg−1 mg kg−1 d−1 mg m−2 d−1
Indicates extraction by 2 M KCl.
C kg−1 , 11 ± 0.10 g N kg−1 , 676 ± 6.85 mg P kg−1 , 32 ± 0.31%, respectively) compared to the bayou sediment (105 ± 8.69 g C kg−1 , 7.0 ± 0.66 g N kg−1 and 589 ± 12.2 mg P kg−1 , 22 ± 0.53% respectively). Extractable NH4 -N was approximately 2.3 times higher in the marsh soil than the bayou sediment at 330 ± 7.17 and 145 ± 8.58 mg N kg−1 , respectively. The PMN rate was 9.63 mg N kg−1 d−1 for the bayou sediment and 9.42 ± 1.62 mg N kg−1 d−1 for the marsh soil. In the marsh soil, TC was negatively correlated with % moisture (r = −0.98, n = 5; Table 2). Extractable NH4 + increased as TP increased (r = 0.96) in marsh soils. Redox potential, measured at 5 cm soil depth, was significantly different between bayou sediment and marsh soils at −240 ± 3.4 and −190 ± 31 mV, respectively (p < 0.01). The pH, salinity, and DO, all measured in the water column, were not significantly different between bayou sediment and marsh soil and therefore were averaged together for 7.32 ± 0.11, 0.07 ± 0.09 ppt and 2.97 ± 0.59 mg L−1 , respectively. 3.2. Ammonification and denitrification rates Water column ammonium in the bayou sediment treatment increased at a rate of 0.121 ± 0.008 mg NH4 -N d−1 (Fig. 2a). Water column ammonium increased at a rate of 0.188 ± 0.033 mg NH4 -N d−1 (Fig. 2b) for the marsh soil treatment. The rate of ammonification was significantly lower for the bayou versus marsh (Table 1). On an areal basis, the ammonification rates were 14.8 ± 0.9 and 23.0 ± 4.1 mg NH4 -N m−2 d−1 , respectively for the bayou sediment and marsh soil. Mean nitrate reduction for the bayou sediment treatment was −0.078 ± 0.012 mg NO3 -N d−1 (Fig. 2a). For the marsh soil treatment, nitrate decreased at a rate of −0.059 ± 0.007 mg NO3 -N d−1 . The rate of nitrate removal was significantly greater for bayou sediment versus marsh soil (Table 1; p = 0.04) and was 9.5 ± 1.5 and 7.2 ± 0.9 mg NO3 -N m−2 d−1 , respectively for bayou sediment and marsh soil on an areal basis. We statistically compared the ammonification rates of both control and treatment (added nitrate) cores for both the bayou and marsh treatments in order to determine if the major nitrate reduction pathway was conservative, as in the case for DNRA, or if N was lost from the system by gaseous removal, as for denitrification. If DNRA was a major N transformation mechanism, then the ammonification rates for the nitrate addition treatments would be higher than that of the controls. For the bayou sediment, the rates of ammonium production were not significantly different between control and treatment cores at 14.1 ± 3.3 and 14.8 ± 0.9 mg NH4 N d−1 , respectively. The rates of ammonification were significantly different between control and treatment cores for the marsh soil (34.0 ± 3.6 and 23.0 ± 4.1 mg NH4 -N d−1 ; p = 0.003); however, the
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Table 2 Product-moment correlation coefficients for bayou sediment and marsh soil characteristics. Bold indicates significance at P < 0.05. Bayou sediment TC TN TP LOI MBC MBN NH4 + PMN
% Moisture 0.00 −0.54 −0.62 0.65 −0.36 −0.25 −0.01 0.96
TC
TN
TP
LOI
MBC
MBN
NH4 +
0.79 −0.27 −0.41 0.57 0.13 0.63 0.95
0.01 −0.70 0.83 0.30 0.43 −1.00
−0.65 −0.04 −0.54 −0.49 −0.89
−0.71 0.37 0.22 0.72
0.03 −0.04 −0.69
0.71 −0.04
0.58
Marsh soil
% Moisture
TC
TN
TP
LOI
MBC
MBN
NH4 +
TC TN TP LOI MBC MBN NH4 + PMNa
−0.98 0.23 0.64 −0.64 0.40 −0.13 0.65 0.39
−0.35 −0.48 0.76 −0.43 0.28 −0.53 0.02
−0.06 −0.24 0.54 0.01 0.22 −0.63
0.17 0.43 0.58 0.96 1.00
−0.08 0.80 0.15 0.67
0.03 0.53 0.95
0.63 0.74
0.98
a
n = 3 for PMN, all other components n = 5, r = 0.88.
ammonification rate was higher in the control cores than the treatment cores. Therefore, our data suggest that DNRA was not a major pathway for nitrate reduction in our system. The rates of ammonium production can likely be attributed to N mineralization while the rates of nitrate reduction likely are through denitrification. Soil properties in the bayou sediment and the marsh soil were significantly different because marsh soils were higher in soil organic matter. Extractable NH4 -N was also higher in marsh soil as organic matter serves as the substrate for N mineralization (White and Reddy, 2000). Both bayou sediment and marsh soil treatments experienced a net loss of nitrate over the 9 d study. The rate of nitrate reduction
a
3.0 Nitrate
Ammonium
2.5
mg N L-1
2.0 1.5 1.0 0.5 0.0
b
0
1
2
3
0
1
2
3
4
5
6
7
8
9
3.0 2.5
mg N L-1
2.0 1.5 1.0 0.5 0.0 4
5
6
7
8
9
Time (days) Fig. 2. Mean water column N concentration over 9 d in a bayou sediment (a) and marsh soil (b), presented as mean ± one sd (n = 4).
was significantly higher in the bayou treatment contrary to our original hypothesis. The higher ammonification rates in the marsh treatment may explain the lower rate of nitrate loss in the marsh soil cores, as the mineralized N could have added greater N to the nitrate pool through nitrification (Seitzinger, 1994; Patrick et al., 1996). Significantly higher TC and extractable NH4 -N in marsh soils support N mineralization in the marsh soils. Increasing water column ammonium concentrations also support higher N mineralization in marsh soil (23.0 ± 4.1 mg NH4 -N m−2 d−1 ) than the bayou sediment (14.8 ± 0.9 mg NH4 -N m−2 d−1 ). Adjacent zones of aerobic and anaerobic conditions are needed for coupling of nitrification and denitrification to occur (White and Reddy, 2003). Nitrification of ammonium to nitrate occurs only under aerobic conditions. Water column DO levels of 2.97 ± 0.59 mg L−1 suggest conditions for nitrification to occur were present. Conversely, denitrification only occurs under anaerobic conditions. Redox measurements of −240 ± 3.4 and −190 ± 31 mV at 5 cm below the sediment/soil surface suggest conditions existed for denitrification to occur (Patrick et al., 1996). Diffusion of ammonium from N mineralization in sediment/soil into the oxygenated water column increases nitrate concentration in the water column (Reddy et al., 1978). In addition, bayou sediments may have had a higher initial active microbial consortium as the river water with associated nitrate was present only in the bayous while the marsh was not flooded by river water at the time of collection. Research has shown that wetland soils/sediments previously in contact with nitrate have a higher active denitrifying enzyme activity (White and Reddy, 1999) and that in general, the higher the concentration of nitrate, the higher the initial enzyme activity (Gardner and White, 2010). The marsh denitrification rates are low, a possible artifact of root removal. Consequently, we compared the rate of nitrate loss for the marsh soil (this study) to a published companion study where the rate of nitrate loss was monitored in a marsh soil colonized by macrophytes at the same site. Marsh soil planted with macrophytes reduced a 2 mg NO3 -N L−1 water column to below detection within 12 h, as opposed to this study, where the marsh soil without plants only reduced about half of the 2 mg NO3 -N L−1 water column over 9 d (VanZomeren et al., 2012). The previously published study partitioned the 15 N-NO3 into soil, plant and gaseous loss components. The rates of nitrate loss attributed to denitrification in that study was over an order of magnitude greater than the denitrification rates seen in this study.
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This comparison clearly demonstrates macrophytes critical role in promoting denitrification (Smith and Tiedje, 1979b). In addition, there was no ammonium release from the macrophyte treatment compared to significant ammonium released in our study with no plants. Consequently, the presence of macrophytes not only increased denitrification but also prevented the release of ammonium due to immobilization. The management implications for this study relate to the operation of the large-scale Mississippi River diversions. Although Mississippi River diversions were built to reduce salinity and enhance wildlife and fisheries in receiving wetlands, it has become clear that these diversions can also help mitigate nutrient loads, particularly nitrate, in diverted Mississippi River water before discharge into the Gulf of Mexico. This, however, can only be achieved if discharge rates are managed to enhance denitrification potential in the receiving basin. At low flow, the high nitrate river water passes through the bayou canals with low denitrification potential and therefore delivers nitrate out to the coastal ocean. At higher flow, the river water floods the vegetated marsh areas which increases the removal rate of nitrate and reduces nitrate export to the coastal ocean. Concerns over possible reductions in marsh stability due to soil C oxidation through denitrification were not supported by stoichiometric calculations in the companion study (VanZomeren et al., 2012), further supporting the use of higher diversion flows to enhance denitrification potential and removal of nitrate before discharging to the coastal ocean. 4. Conclusion Net nitrate removal was significantly higher in bayou sediments compared with marsh soil when roots are removed (9.5 ± 1.5 and 7.2 ± 0.9 mg NO3 -N m−2 d−1 , respectively). The ammonification rate is significantly lower in bayou sediments than marsh soil (14.8 ± 0.9 and 23.0 ± 4.1 mg NH4 -N m−2 d−1 , respectively). This ammonium production suggests that both bayou sediments and marsh soil can potentially be a source of nitrate through coupled N mineralization and nitrification, thereby decreasing potential nitrate removal capacity by denitrification. Comparison of our rates with a companion study that included plants suggest nitrate reduction rates in vegetated and non-vegetated marsh soils indicated that the rate of denitrification increased tenfold in vegetated soils as well as assimilation of released ammonium into biomass. Management of diversion flow rate is critical for enhancing the potential reduction of nitrate by soil denitrification. Lower flows do not allow the vegetated marsh to be flooded, resulting in a lower denitrification potential and less nitrate removal. Therefore, the operation of river diversions should be maintained on the high flow end of the spectrum to assure that nitrate is delivered into the vegetated marsh, if the goal is maximum nitrate removal from river water. Acknowledgments We would like to acknowledge Louisiana Sea Grant Program and the Louisiana Office of Coastal Protection and Restoration for providing funding for this research by way of a Coastal Sciences Assistantship. We also would like to thank Nathan Nguyen and Anthony Nguyen for laboratory assistance.
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