Human activities have led to 1–2% of coastal wetlands lost per year globally, with subsequent losses in ecosystem services such as nutrient filtering and carbon sequestration. Wetland construction is used to mitigate losses of marsh cover and services resulting from human impacts in coastal areas. Though marsh structure can recover relatively quickly (i.e., <10 years) after construction, there are often long‐term lags in the recovery of ecosystem functions in constructed marshes. We conducted a year‐long study comparing seasonal plant productivity, ecosystem respiration (
Denitrification and dissimilatory nitrate reduction to ammonium (DNRA) both require low oxygen and high organic carbon conditions common in wetland ecosystems. Denitrification permanently removes nitrogen from the ecosystem as a gas while DNRA recycles nitrogen within the ecosystem via production of ammonium. The relative prevalence of denitrification versus DNRA has implications for the fate of nitrate in ecosystems. Unplanned and unmanaged urban accidental wetlands in the Salt River channel near downtown Phoenix, Arizona, USA receive high nitrate relative to non‐urban wetlands and have a high capacity for denitrification, but unknown capacity for DNRA. We conducted in‐situ push‐pull tests with isotopically labeled nitrate to measure denitrification and DNRA rates in three of the dominant vegetative patch types in these urban accidental wetlands. DNRA accounted for between 2% and 40% of nitrate reduction (DNRA plus denitrification) with the highest rates measured in patches of
- Award ID(s):
- 1832016
- NSF-PAR ID:
- 10363391
- Publisher / Repository:
- DOI PREFIX: 10.1029
- Date Published:
- Journal Name:
- Journal of Geophysical Research: Biogeosciences
- Volume:
- 127
- Issue:
- 2
- ISSN:
- 2169-8953
- Format(s):
- Medium: X
- Sponsoring Org:
- National Science Foundation
More Like this
-
more » « less
), denitrification, and dissimilatory nitrate reduction to ammonium (DNRA) between two 33‐year‐old constructed marshes (CON‐1, CON‐2) and a nearby natural reference marsh (NAT). We found that CON‐1 and CON‐2 were structurally similar to NAT (i.e., plant aboveground and belowground biomass did not differ). Likewise, gross ecosystem productivity (GEP), , and net ecosystem exchange (NEE) were similar across all marshes. Further, DNRA and denitrification were similar across marshes, with the exception of greater denitrification rates at CON‐2 than at the other two sites. While pore‐water ammonium concentrations were similar across all marshes, organic matter (OM) content, pore‐water phosphate, nitrate + nitrite, and hydrogen sulfide concentrations were greater in NAT than CON‐1 and CON‐2. Collectively, this work suggests that current marsh construction practices could be a suitable tool for recovering plant structure and some ecosystem functions. However, the lag in recovery of pore‐water nutrient stocks and OM content also suggests that some biogeochemical functions may take longer than a few decades to fully recover in constructed marshes. -
more » « less
Abstract Coastal marshes mitigate allochthonous nitrogen (N) inputs to adjacent marine habitat; however, their extent is declining rapidly. As a result, marsh restoration and construction have become a major foci of wetland management. Constructed marshes can quickly reach similar plant biomass to natural marshes, but biogeochemical functions like N removal and retention can take decades to reach functional equivalency, often due to lags in organic matter (OM) pools development in newly constructed marshes. We compared denitrification and dissimilatory nitrate reduction to ammonium (DNRA) rates in a 32 year-old constructed marsh and adjacent reference marsh in the Northern Gulf of Mexico. Marsh sediments packed into 3 mm “thin discs” were subjected to three OM quality treatments (no OM addition, labile OM, or recalcitrant OM) and two N treatments (ambient nitrate or elevated nitrate) during a 13 day incubation. We found that OM addition, rather than marsh type or nitrate treatment, was the most important driver of nitrate reduction, increasing both denitrification and DNRA and promoting DNRA over denitrification in both marshes. Fungal and bacterial biomass were higher in the natural marsh across treatments, but recalcitrant OM increased fungal biomass in the constructed marsh, suggesting OM-limitation of fungal growth. We found that constructed marshes are capable of similar denitrification and DNRA as natural marshes after 30 years, and that labile OM addition promotes N retention in both natural and constructed marshes.
Graphical Abstract Conceptual figure highlighting the findings of this experiment. Under control treatment with no C addition (bottom panel), constructed and natural marshes have similar rates of both DNRA and denitrification. The natural marsh has higher fungal and bacterial biomass, while fungal biomass is not detectable in the constructed marsh. Under labile OM additions (upper left panel), rates of both DNRA and denitrification are increased and DNRA becomes favored over denitrification in both marshes. Recalcitrant OM additions (upper right) increase denitrification, but do not affect DNRA or % denitrification. The addition of recalcitrant OM also increases the detectability of fungal biomass in the constructed marsh.
-
null (Ed.)Understanding the dominant soil nitrogen (N) cycling processes in southern Appalachian forests is crucial for predicting ecosystem responses to changing N deposition and climate. The role of anaerobic nitrogen cycling processes in well-aerated soils has long been questioned, and recent N cycling research suggests it needs to be re-evaluated. We assessed gross and potential rates of soil N cycling processes, including mineralization, nitrification, denitrification, nitrifier denitrification, and dissimilatory nitrate reduction to ammonium (DNRA) in sites representing a vegetation and elevation gradient in the U.S. Department of Agriculture (USDA) Forest Service Experimental Forest, Coweeta Hydrologic Laboratory in southwestern North Carolina, USA. N cycling processes varied among sites, with gross mineralization and nitrification being greatest in high-elevation northern hardwood forests. Gaseous N losses via nitrifier denitrification were common in all ecosystems but were greatest in northern hardwood. Ecosystem N retention via DNRA (nitrification-produced NO3 reduced to NH4) ranged from 2% to 20% of the total nitrification and was highest in the mixed-oak forest. Our results suggest the potential for gaseous N losses through anaerobic processes (nitrifier denitrification) are prevalent in well-aerated forest soils and may play a key role in ecosystem N cycling.more » « less
-
Site description. This data package consists of data obtained from sampling surface soil (the 0-7.6 cm depth profile) in black mangrove (Avicennia germinans) dominated forest and black needlerush (Juncus roemerianus) saltmarsh along the Gulf of Mexico coastline in peninsular west-central Florida, USA. This location has a subtropical climate with mean daily temperatures ranging from 15.4 °C in January to 27.8 °C in August, and annual precipitation of 1336 mm. Precipitation falls as rain primarily between June and September. Tides are semi-diurnal, with 0.57 m median amplitudes during the year preceding sampling (U.S. NOAA National Ocean Service, Clearwater Beach, Florida, station 8726724). Sea-level rise is 4.0 ± 0.6 mm per year (1973-2020 trend, mean ± 95 % confidence interval, NOAA NOS Clearwater Beach station). The A. germinans mangrove zone is either adjacent to water or fringed on the seaward side by a narrow band of red mangrove (Rhizophora mangle). A near-monoculture of J. roemerianus is often adjacent to and immediately landward of the A. germinans zone. The transition from the mangrove to the J. roemerianus zone is variable in our study area. An abrupt edge between closed-canopy mangrove and J. roemerianus monoculture may extend for up to several hundred meters in some locations, while other stretches of ecotone present a gradual transition where smaller, widely spaced trees are interspersed into the herbaceous marsh. Juncus roemerianus then extends landward to a high marsh patchwork of succulent halophytes (including Salicornia bigellovi, Sesuvium sp., and Batis maritima), scattered dwarf mangrove, and salt pans, followed in turn by upland vegetation that includes Pinus sp. and Serenoa repens. Field design and sample collection. We established three study sites spaced at approximately 5 km intervals along the western coastline of the central Florida peninsula. The sites consisted of the Salt Springs (28.3298°, -82.7274°), Energy Marine Center (28.2903°, -82.7278°), and Green Key (28.2530°, -82.7496°) sites on the Gulf of Mexico coastline in Pasco County, Florida, USA. At each site, we established three plot pairs, each consisting of one saltmarsh plot and one mangrove plot. Plots were 50 m^2 in size. Plots pairs within a site were separated by 230-1070 m, and the mangrove and saltmarsh plots composing a pair were 70-170 m apart. All plot pairs consisted of directly adjacent patches of mangrove forest and J. roemerianus saltmarsh, with the mangrove forests exhibiting a closed canopy and a tree architecture (height 4-6 m, crown width 1.5-3 m). Mangrove plots were located at approximately the midpoint between the seaward edge (water-mangrove interface) and landward edge (mangrove-marsh interface) of the mangrove zone. Saltmarsh plots were located 20-25 m away from any mangrove trees and into the J. roemerianus zone (i.e., landward from the mangrove-marsh interface). Plot pairs were coarsely similar in geomorphic setting, as all were located on the Gulf of Mexico coastline, rather than within major sheltering formations like Tampa Bay, and all plot pairs fit the tide-dominated domain of the Woodroffe classification (Woodroffe, 2002, "Coasts: Form, Process and Evolution", Cambridge University Press), given their conspicuous semi-diurnal tides. There was nevertheless some geomorphic variation, as some plot pairs were directly open to the Gulf of Mexico while others sat behind keys and spits or along small tidal creeks. Our use of a plot-pair approach is intended to control for this geomorphic variation. Plot center elevations (cm above mean sea level, NAVD 88) were estimated by overlaying the plot locations determined with a global positioning system (Garmin GPS 60, Olathe, KS, USA) on a LiDAR-derived bare-earth digital elevation model (Dewberry, Inc., 2019). The digital elevation model had a vertical accuracy of ± 10 cm (95 % CI) and a horizontal accuracy of ± 116 cm (95 % CI). Soil samples were collected via coring at low tide in June 2011. From each plot, we collected a composite soil sample consisting of three discrete 5.1 cm diameter soil cores taken at equidistant points to 7.6 cm depth. Cores were taken by tapping a sleeve into the soil until its top was flush with the soil surface, sliding a hand under the core, and lifting it up. Cores were then capped and transferred on ice to our laboratory at the University of South Florida (Tampa, Florida, USA), where they were combined in plastic zipper bags, and homogenized by hand into plot-level composite samples on the day they were collected. A damp soil subsample was immediately taken from each composite sample to initiate 1 y incubations for determination of active C and N (see below). The remainder of each composite sample was then placed in a drying oven (60 °C) for 1 week with frequent mixing of the soil to prevent aggregation and liberate water. Organic wetland soils are sometimes dried at 70 °C, however high drying temperatures can volatilize non-water liquids and oxidize and decompose organic matter, so 50 °C is also a common drying temperature for organic soils (Gardner 1986, "Methods of Soil Analysis: Part 1", Soil Science Society of America); we accordingly chose 60 °C as a compromise between sufficient water removal and avoidance of non-water mass loss. Bulk density was determined as soil dry mass per core volume (adding back the dry mass equivalent of the damp subsample removed prior to drying). Dried subsamples were obtained for determination of soil organic matter (SOM), mineral texture composition, and extractable and total carbon (C) and nitrogen (N) within the following week. Sample analyses. A dried subsample was apportioned from each composite sample to determine SOM as mass loss on ignition at 550 °C for 4 h. After organic matter was removed from soil via ignition, mineral particle size composition was determined using a combination of wet sieving and density separation in 49 mM (3 %) sodium hexametaphosphate ((NaPO_3)_6) following procedures in Kettler et al. (2001, Soil Science Society of America Journal 65, 849-852). The percentage of dry soil mass composed of silt and clay particles (hereafter, fines) was calculated as the mass lost from dispersed mineral soil after sieving (0.053 mm mesh sieve). Fines could have been slightly underestimated if any clay particles were burned off during the preceding ignition of soil. An additional subsample was taken from each composite sample to determine extractable N and organic C concentrations via 0.5 M potassium sulfate (K_2SO_4) extractions. We combined soil and extractant (ratio of 1 g dry soil:5 mL extractant) in plastic bottles, reciprocally shook the slurry for 1 h at 120 rpm, and then gravity filtered it through Fisher G6 (1.6 μm pore size) glass fiber filters, followed by colorimetric detection of nitrite (NO_2^-) + nitrate (NO_3^-) and ammonium (NH_4^+) in the filtrate (Hood Nowotny et al., 2010,Soil Science Society of America Journal 74, 1018-1027) using a microplate spectrophotometer (Biotek Epoch, Winooski, VT, USA). Filtrate was also analyzed for dissolved organic C (referred to hereafter as extractable organic C) and total dissolved N via combustion and oxidation followed by detection of the evolved CO_2 and N oxide gases on a Formacs HT TOC/TN analyzer (Skalar, Breda, The Netherlands). Extractable organic N was then computed as total dissolved N in filtrate minus extractable mineral N (itself the sum of extractable NH_4-N and NO_2-N + NO_3-N). We determined soil total C and N from dried, milled subsamples subjected to elemental analysis (ECS 4010, Costech, Inc., Valencia, CA, USA) at the University of South Florida Stable Isotope Laboratory. Median concentration of inorganic C in unvegetated surface soil at our sites is 0.5 % of soil mass (Anderson, 2019, Univ. of South Florida M.S. thesis via methods in Wang et al., 2011, Environmental Monitoring and Assessment 174, 241-257). Inorganic C concentrations are likely even lower in our samples from under vegetation, where organic matter would dilute the contribution of inorganic C to soil mass. Nevertheless, the presence of a small inorganic C pool in our soils may be counted in the total C values we report. Extractable organic C is necessarily of organic C origin given the method (sparging with HCl) used in detection. Active C and N represent the fractions of organic C and N that are mineralizable by soil microorganisms under aerobic conditions in long-term soil incubations. To quantify active C and N, 60 g of field-moist soil were apportioned from each composite sample, placed in a filtration apparatus, and incubated in the dark at 25 °C and field capacity moisture for 365 d (as in Lewis et al., 2014, Ecosphere 5, art59). Moisture levels were maintained by frequently weighing incubated soil and wetting them up to target mass. Daily CO_2 flux was quantified on 29 occasions at 0.5-3 week intervals during the incubation period (with shorter intervals earlier in the incubation), and these per day flux rates were integrated over the 365 d period to compute an estimate of active C. Observations of per day flux were made by sealing samples overnight in airtight chambers fitted with septa and quantifying headspace CO_2 accumulation by injecting headspace samples (obtained through the septa via needle and syringe) into an infrared gas analyzer (PP Systems EGM 4, Amesbury, MA, USA). To estimate active N, each incubated sample was leached with a C and N free, 35 psu solution containing micronutrients (Nadelhoffer, 1990, Soil Science Society of America Journal 54, 411-415) on 19 occasions at increasing 1-6 week intervals during the 365 d incubation, and then extracted in 0.5 M K_2SO_4 at the end of the incubation in order to remove any residual mineral N. Active N was then quantified as the total mass of mineral N leached and extracted. Mineral N in leached and extracted solutions was detected as NH_4-N and NO_2-N + NO_3-N via colorimetry as above. This incubation technique precludes new C and N inputs and persistently leaches mineral N, forcing microorganisms to meet demand by mineralizing existing pools, and thereby directly assays the potential activity of soil organic C and N pools present at the time of soil sampling. Because this analysis commences with disrupting soil physical structure, it is biased toward higher estimates of active fractions. Calculations. Non-mobile C and N fractions were computed as total C and N concentrations minus the extractable and active fractions of each element. This data package reports surface-soil constituents (moisture, fines, SOM, and C and N pools and fractions) in both gravimetric units (mass constituent / mass soil) and areal units (mass constituent / soil surface area integrated through 7.6 cm soil depth, the depth of sampling). Areal concentrations were computed as X × D × 7.6, where X is the gravimetric concentration of a soil constituent, D is soil bulk density (g dry soil / cm^3), and 7.6 is the sampling depth in cm.more » « less
-
more » « less
Human activities have decreased global salt marsh surface area with a subsequent loss in the ecosystem functions they provide. The creation of marshes in terrestrial systems has been used to mitigate this loss in marsh cover. Although these constructed marshes may rapidly recover ecosystem structure, biogeochemical processes may be slow to recover. We compared denitrification and dissimilatory nitrate reduction to ammonium (DNRA) rates between a 32‐year‐old excavation‐created salt marsh (CON‐2) and a nearby natural reference salt marsh (NAT) to assess the recovery of ecosystem function. These process rates were measured at 5 cm increments to a depth of 25 cm to assess how plant rooting depth and organic matter accumulation impact N‐cycling. We found that, for both marshes, denitrification and DNRA declined with depth with the highest rates occurring in the top 10 cm. In both systems, N‐retention by DNRA accounted for upwards of 75% of nitrate reduction, but denitrification and DNRA rates were nearly 2× and 3× higher in NAT than CON‐2, respectively. Organic matter was 6× lower in CON‐2, likely due to limited plant belowground biomass production. However, there was no response to glucose additions, suggesting that the microbial functional community, not substrate limitation, limited nitrate reduction recovery. Response ratios showed that denitrification in CON‐2 recovered in surficial sediments where belowground biomass was highest, even though biomass recovery was minimal. This indicates that although recovery of ecosystem function was constrained, it occurred on a faster trajectory than that of ecosystem structure.
