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Received Date : 07-Oct-2014 Accepted Date : 02-Feb-2015 Article type : Research Review Title: Nitrous oxide fluxes in estuarine environments: Response to global change Running head: Nitrous oxide in estuarine environments
Authors: Rachel H. Murray, Phone: +61 0478789470, Email:
[email protected] Dirk V. Erler, Phone: +61 0266203256, Email:
[email protected] Bradley D. Eyre, Phone: +61 0266203773, Email:
[email protected] Institute of origin (all authors can be contacted at this address): Centre for Coastal Biogeochemistry, Southern Cross University, PO Box 157, Lismore, NSW 2480, Australia
Corresponding author: Rachel Murray, Southern Cross University, PO Box 157, Lismore, NSW 2480, Australia Phone: +61 0478789470 Email:
[email protected] Keywords: nitrous oxide, mangrove, salt marsh, estuary, mudflat, greenhouse gas,
intertidal, denitrification
Abstract Nitrous oxide is a powerful, long-lived greenhouse gas, but we know little about the role
of estuarine areas in the global N2O budget. This review summarizes 56 studies of N2O fluxes and associated biogeochemical controlling factors in estuarine open waters, salt marshes, mangroves, and intertidal sediments. The majority of in situ N2O production occurs as a result of This article has been accepted for publication and undergone full peer review but has not been through the copyediting, typesetting, pagination and proofreading process, which may lead to differences between this version and the Version of Record. Please cite this article as doi: 10.1111/gcb.12923 This article is protected by copyright. All rights reserved.
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sediment denitrification, although the water column contributes N2O through nitrification in suspended particles. The most important factors controlling N2O fluxes seem to be dissolved
inorganic nitrogen (DIN) and oxygen availability, which in turn are affected by tidal cycles, groundwater inputs, and macrophyte density. The heterogeneity of coastal environments leads to a high variability in observations, but on average estuarine open water, intertidal and vegetated environments are sites of a small positive N2O flux to the atmosphere (range 0.15–0.91; median 0.31; Tg N2O-N yr-1). Global changes in macrophyte distribution and anthropogenic nitrogen loading are expected to increase N2O emissions from estuaries. We estimate that a doubling of
current median NO3- concentrations would increase the global estuary water-air N2O flux by about 0.45 Tg N2O-N yr-1 or about 190%. A loss of 50% of mangrove habitat, being converted to unvegetated intertidal area, would result in a net decrease in N2O emissions of 0.002 Tg N2O-N
yr-1. In contrast, conversion of 50% of salt marsh to unvegetated area would result in a net increase of 0.001 Tg N2O-N yr-1. Decreased oxygen concentrations may inhibit production of N2O by nitrification, however sediment denitrification and the associated ratio of N2O:N2 is
expected to increase.
Introduction Nitrous oxide is a powerful, long-lived greenhouse gas [Montzka et al., 2011], and
contributes to ozone destruction in the stratosphere [Portmann et al., 2012]. The concentration of N2O in the atmosphere is increasing by about 0.25% yr-1 [Stocker et al., 2013], largely attributed to human activity such as agricultural fertilization, sewage release, and industrial waste discharge [Smith et al., 1997] . In recent decades many estuaries have been greatly altered by increased nitrogen (N)
loads from wastewater and agricultural runoff [Galloway et al., 2003; Howarth et al., 1996]. In sediments and waterways affected by increased N, microbes and primary producers use this excess N and produce a variety of byproducts, including N2O gas [Poth and Focht, 1985; Wrage et al., 2001]. Estuaries and coastal wetlands are sites of intense biological production [Alongi et al., 1998], and likely play an important role in the global N2O budget. However, little is known about the magnitude of the N2O flux from certain coastal environments. While estimates exist for global N2O fluxes from mangroves and estuarine systems as a whole, there have been no global
estimates of N2O emission from intertidal bare sediments, salt marshes, or seagrasses.
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from N2O produced in-situ, especially where tides and currents are strong [Bange, 2006; Leip, 2000]. While less affected by allochthonous N2O, calmer intertidal and vegetated areas can
exhibit complex tidal variation. Where NO3- is scarce, or where NH4+ dominates the DIN pool, N2O emissions may increase after sediment exposure, likely due to increased sediment aeration and degassing of N2O [Cheng et al., 2007; Wang et al., 2007]. The effect of sediment exposure
can be quite strong; At Nakaumi Lake, for example, the variability in N2O flux due to tidal
position outweighs diurnal and temporal variability [Hirota et al., 2007]. However, at field sites with high natural or anthropogenic nitrate concentrations, the water column is the primary source of nitrate for sedimentary denitrifiers, [e.g. Adams et al., 2012; Dong et al., 2006; Dong et al., 2002; Robinson et al., 1998; Tong et al., 2013] especially in highly bioturbated sediments [Law
et al., 1991]. Aerial exposure removes this nitrate source, slowing N2O production and emission, in particular in salt marshes and intertidal mudflats [Adams et al., 2012; Robinson et al., 1998].
As mentioned previously, N2O concentrations are typically greatest at upstream locations
within the estuary [Bange et al., 1996; Zhang et al., 2010]. However, in many estuaries the spatial pattern reflects a combination of environmental factors. For example, the Adyar River, in India, has a mid-estuarine N2O maximum [Rajkumar et al., 2008], and many European estuaries exhibit a spike in N2O emissions within a narrow salinity band, near the turbidity maximum
[Abril et al., 2000; Bange, 2006; Barnes and Owens, 1999; De Bie et al., 2002; De Wilde and De
Bie, 2000]. In these cases, N2O efflux responds to the residence time and movement of DIN through the catchment, as well as the concentration of suspended particulates. Other studies note the presence of N2O “hot spots.” [Corredor et al., 1999; Marty et al., 2001; Wong et al., 2013] or
find little correlation between N2O and salinity [Bange et al., 1998; Zhang et al., 2010]. In some cases, groundwater can be a confounding factor. DIN and N2O concentrations
are usually much higher in groundwater than in receiving water [Burnett and Dulaiova, 2003] and groundwater can be a significant N2O source into estuaries [LaMontagne et al., 2003; Wong
et al., 2013]. Tidal pumping of groundwater can also contribute to tidal variability in N2O
concentrations in the water column [Wong et al., 2013]. Groundwater N2O concentrations increase with increasing subterranean denitrification in nearby aquifers [LaMontagne et al.,
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Betlach, M. R., and J. M. Tiedje (1981), Kinetic explanation for accumulation of nitrite, nitric oxide, and nitrous oxide during bacterial denitrification, Applied and Environmental Microbiology, 42(6), 1074-1084. Binnerup, S. J., K. Jensen, N. P. Revsbech, M. H. Jensen, and J. Sørensen (1992), Denitrification, dissimilatory reduction of nitrate to ammonium, and nitrification in a bioturbated estuarine sediment as measured with 15N and microsensor techniques, Applied and environmental microbiology, 58(1), 303-313. Biswas, H., A. Chatterjee, S. K. Mukhopadhya, T. K. De, S. Sen, and T. K. Jana (2005), Estimation of ammonia exchange at the land–ocean boundary condition of Sundarban mangrove, northeast coast of Bay of Bengal, India, Atmospheric Environment, 39(25), 4489-4499. Blackburn, T. H., D. B. Nedwell, and W. J. Wiebe (1994), Active mineral cycling in a Jamaican seagrass sediment, Marine Ecology Progress Series, 110(2-3), 233-239. Blackwell, M. S. A., S. Yamulki, and R. Bol (2010), Nitrous oxide production and denitrification rates in estuarine intertidal saltmarsh and managed realignment zones, Estuarine, Coastal and Shelf Science, 87(4), 591-600. Borges, A. V., and G. Abril (2011), Carbon Dioxide and Methane Dynamics in Estuaries, 119161 pp. Burgin, A. J., J. G. Lazar, P. M. Groffman, A. J. Gold, and D. Q. Kellogg (2013), Balancing nitrogen retention ecosystem services and greenhouse gas disservices at the landscape scale, Ecological Engineering, 56, 26-35. Burnett, W. C., and H. Dulaiova (2003), Estimating the dynamics of groundwater input into the coastal zone via continuous radon-222 measurements, Journal of environmental radioactivity, 69(1), 21-35.
Butler, J. H., J. W. Elkins, T. M. Thompson, and K. B. Egan (1989), Tropospheric and dissolved N2O of the west Pacific and east Indian Oceans during the El Nino Southern Oscillation event of 1987, Journal of Geophysical Research-Atmospheres, 94(D12), 14865-14877. Chang, C., H. H. Janzen, C. M. Cho, and E. M. Nakonechny (1998), Nitrous oxide emission through plants, Soil Science Society of America Journal, 62(1), 35-38. Chapuis-Lardy, L., N. Wrage, A. Metay, J.-L. Chotte, and M. Bernoux (2007), Soils, a sink for N2O? A review, Global Change Biology, 13(1), 1-17.
Chauhan, R., A. L. Ramanathan, and T. K. Adhya (2008), Assessment of methane and nitrous oxide flux from mangroves along Eastern coast of India, Geofluids, 8(4), 321-332.
Chen, G. C., N. F. Y. Tam, and Y. Ye (2010), Summer fluxes of atmospheric greenhouse gases N2O, CH4 and CO2 from mangrove soil in South China, Science of the Total Environment, 408(13), 2761-2767.
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Welsh, D., M. Bartoli, D. Nizzoli, G. Castaldelli, S. A. Riou, and P. Viaroli (2000), Denitrification, nitrogen fixation, community primary productivity and inorganic-N and oxygen fluxes in an intertidal Zostera noltii meadow, Marine Ecology Progress Series, 208(5).
Welsh, D., G. Castadelli, M. Bartoli, D. Poli, M. Careri, R. de Wit, and P. Viaroli (2001), Denitrification in an intertidal seagrass meadow, a comparison of 15N-isotope and acetyleneblock techniques: dissimilatory nitrate reduction to ammonia as a source of N2O?, Marine Biology, 139(6), 1029-1036. Whigham, D. F., J. T. a. Verhoeven, V. Samarkin, and P. J. Megonigal (2009), Responses of Avicennia germinans (Black Mangrove) and the Soil Microbial Community to Nitrogen Addition in a Hypersaline Wetland, Estuaries and Coasts, 32(5), 926-936. Wong, W. W., M. R. Grace, I. Cartwright, M. B. Cardenas, P. B. Zamora, and P. L. M. Cook (2013), Dynamics of groundwater-derived nitrate and nitrous oxide in a tidal estuary from radon mass balance modeling, Limnology and Oceanography, 58(5), 1-19. Woodwell, G. (1980), Aquatic systems as part of the biosphere, Fundamentals of aquatic ecosystems. Blackwell, Oxford, 201-215. Wrage, N., G. Velthof, M. Van Beusichem, and O. Oenema (2001), Role of nitrifier denitrification in the production of nitrous oxide, Soil Biology and Biochemistry, 33(12), 17231732.
Yan, W., L. Yang, F. Wang, J. Wang, and P. Ma (2012), Riverine N2O concentrations, exports to estuary and emissions to atmosphere from the Changjiang River in response to increasing nitrogen loads, Global Biogeochemical Cycles, 26(4). Ye, Y., C. Lu, and P. Lin (2000), Seasonal and spatial changes of methane emissions from mangrove wetlands in Hainan and Xiamen, Chinese Journal of Atmospheric Science, 24, 152156. Yu, K., and G. Chen (2009), Nitrous oxide emissions from terrestrial plants: observations, mechanisms and implications, in Nitrous Oxide Emissions Research Progress, edited by B. E. Sheldon AI, pp. 85-104, Nova Science Publishers, Inc, New York, USA. Yu, Z., Y. Li, H. Deng, D. Wang, Z. Chen, and S. Xu (2012), Effect of Scirpus mariqueter on nitrous oxide emissions from a subtropical monsoon estuarine wetland, Journal of Geophysical Research: Biogeosciences, 117(G2), 1-9.
Zhang, G., J. Zhang, J. Xu, and F. Zhang (2006), Distributions, sources and atmospheric fluxes of nitrous oxide in Jiaozhou Bay, Estuarine, Coastal and Shelf Science, 68(3-4), 557-566. Zhang, G., J. Zhang, J. L. Ren, J. B. Li, and S. M. Liu (2008), Distributions and sea-to-air fluxes of methane and nitrous oxide in the North East China Sea in summer, Marine Chemistry, 110(12), 42-55.
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Zhang, G., J. Zhang, S. M. Liu, J. L. Ren, and Y. C. Zhao (2010), Nitrous oxide in the Changjiang (Yangtze River) Estuary and its adjacent marine area: Riverine input, sediment release and atmospheric fluxes, Biogeosciences, 7(11), 3505-3516. Zhao, B., H. Guo, Y. Yan, Q. Wang, and B. Li (2008), A simple waterline approach for tidelands using multi-temporal satellite images: a case study in the Yangtze Delta, Estuarine, Coastal and Shelf Science, 77(1), 134-142. Figure captions Figure 1. Locations of N2O flux measurements worldwide (a), in Europe (b), and in East Asia (c).
Figure 2. Latitudinal distribution of mangrove N2O flux field sites (from this review), and
mangrove areal extent (from Giri et al., 2001). Figure 3. Aqueous NO3- (a) and NH4+ (b) concentrations vs. water-to-air estuarine N2O fluxes. In the Humber River and Adyar River, average N2O fluxes are lower than expected for the given
DIN concentrations. Figure 4. Fluxes of N2O (Tg N2O-N yr-1) under current conditions (a), and under a future scenario (b). The future scenario considers N2O increase due to a doubling of DIN in main
estuary waters (but does not consider DIN changes in other environments) and the change in N2O fluxes due to 50% loss of mangroves, salt marsh, and seagrass.
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Environment
Global Flux Estimate (Tg N2O-N yr-1)
Reference
Mangrove
0.01 - 0.09
This study
Mangrove
0.014 - 0.56
Corredor et al., 1999
Mangrove
0.11
Chauhan et al., 2008
Mangrove
0.07
Barnes et al., 2006
Salt marsh
0.001 - 0.14
This study
Intertidal sediment
0.007 - 0.05
This study
Seagrass
0 - 0.19
This study
Estuarine open water
0.13 - 0.44
This study
Estuarine open water
0.13 - 0.45
Robinson et al., 1998
Estuarine open water
0.22
Law et al., 1992
Estuarine open water
1.5
De Wilde and De Bie (2000)
Estuarine open water
3.7 - 5.7
Bange et al., 1996
Estuarine open water
0.07 - 0.69
Seitzinger and Kroeze, 1998
Estuarine open water Total of estuarine environments Rivers and estuaries combined
0.25
Kroeze et al., 2005
0.17 - 0.95
This study
0.6
Syakila et al., 2010
Table 7. Estimated global N2O fluxes from previous studies.
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consumption by denitrifiers [Chapuis-Lardy et al., 2007]. However, occasionally N2O uptake can occur in the presence of abundant DIN, especially if NH4+ dominates the DIN pool [e.g.
Adams et al., 2012; Daniel et al., 2013; Middelburg et al., 1995; Rajkumar et al., 2008]. Sediment water content and structure affect the rate of nitrous oxide consumption. It is
well understood that high water content impedes gas diffusion through soil and limits sediment aeration, increasing the residence time of N2O in the sediment and providing greater opportunity for N2O consumption by denitrifiers [Letey et al., 1980; Van Groenigen et al., 2005]. This effect
can be seen in vertical profiles of sediment N2O concentration. If the rhizosphere lies underneath a relatively anoxic topsoil layer, the oxygenation of the subsurface leads to N2O production that
is subsequently denitrified to N2 as gases move up through the soil column [Sun et al., 2013]. However, the vertical distribution of both DIN and N2O can vary considerably [Allen et al.,
2007; Bauza et al., 2002; Fernandes et al., 2010; Kristensen et al., 1988]. More thorough sampling could help better characterize the bacterial and physical mechanisms underlying negative N2O fluxes. In both sediments and water, anoxic conditions tend to favor denitrification, while high
oxygen concentrations favor nitrification or combined nitrification-denitrification pathways [Barnes and Owens, 1999; Jensen et al., 1984]. The most intense N2O production occurs under
low oxygen and high-NOx conditions [Barnes and Owens, 1999; Chen et al., 2011; Jensen et al., 1984], in which denitrification activity is stimulated by high DIN and N2O is not scavenged as an electron receptor. These conditions are typical of eutrophic, stagnant waters or the upper reaches of estuaries. Likewise N2O fluxes are high under moderate oxygen and high NH4+ conditions [e.g.
Chen et al., 2010; Corredor et al., 1999; Muñoz-Hincapié et al., 2002], in which nitrification or
coupled nitrification-denitrification can produce abundant N2O. In the case where the dominant DIN compound is NH4+, coupled nitrification-denitrification is likely necessary to produce large
amounts of N2O; nitrification provides the NO3- and NO2- used by denitrifiers, with N2O resulting from either one or both processes.
Oxygen diffusion into sediment is often limited to a few millimeters of depth [Revsbech
et al., 1980], so sediment denitrification usually produces more N2O than sediment nitrification. However, the N2O produced by nitrification is more likely to escape via diffusion before it is
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consumed, so the overall balance between nitrification and denitrification, as a source for N2O, responds to changes in sediment structure [Meyer et al., 2008]. Photosynthesis and gas transport through plant aerenchyma can deliver O2 to the rhizosphere [Colmer, 2003; Jorgensen et al., 2012] potentially suppressing denitrification and leading to lower N2O fluxes in some estuary
sediments [e.g. Moseman-Valtierra et al., 2011; Wang et al., 2007]. Conversely, in some
environments, increased rhizosphere O2 can encourage coupled nitrification-denitrification, leading to increased N2O emissions, provided that NH4+ is available [Risgaard-Petersen et al., 1994; Shieh and Yang, 1997; Usui et al., 2001]. In tidal areas with active benthic fauna, oxygen
may reach further into the sediment profile as a result of bioturbation and bioirrigation [Howes et al., 1986; Kristensen et al., 1985; McKee et al., 1988; Volkenborn et al., 2007], promoting coupled nitrification-denitrification [Banks et al., 2013; Webb and Eyre, 2004]. In addition, benthic fluxes of N2O into the water column are generally higher in bioturbated sediments [Binnerup et al., 1992; Law et al., 1991; Svensson et al., 2001].
In contrast with the sediment, water column oxygen concentrations are usually high and
nitrification is responsible for most N2O production [Barnes and Owens, 1999; Barnes and
Upstill-Goddard, 2011; De Wilde and De Bie, 2000]. This nitrification activity is enhanced when the water has a high sediment load and a high residence time in the estuary [Gonçalves et al., 2010], as the suspended particulates act as a substrate for nitrifying bacteria. Water-column denitrification is usually limited to low oxygen zones, however some denitrification may occur in anoxic microsites, even in relatively oxygenated waters [De Bie et al., 2002]. Analysis of bacteria found in estuarine waters reveals that denitrifiers are more common in low oxygen, low salinity waters, while nitrifiers dominate the saline, oxygenated ocean waters [Marty et al., 2001].
Tidal cycles of flooding and exposure have a strong effect on N2O production and
emission. The water-to-air N2O flux generally increases as the tide drops, because the outgoing
tide carries with it a higher DIN concentration, including anthropogenic N loads introduced upstream [Ferrón et al., 2007]. Exceptions can be found where the major source of DIN is
oceanic, as in upwelling zones [e.g. Daniel et al., 2013]. The outgoing tide may also transport dissolved N2O, and several studies have noted the difficulty in distinguishing allochthonous N2O
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from N2O produced in-situ, especially where tides and currents are strong [Bange, 2006; Leip, 2000]. While less affected by allochthonous N2O, calmer intertidal and vegetated areas can
exhibit complex tidal variation. Where NO3- is scarce, or where NH4+ dominates the DIN pool, N2O emissions may increase after sediment exposure, likely due to increased sediment aeration and degassing of N2O [Cheng et al., 2007; Wang et al., 2007]. The effect of sediment exposure
can be quite strong; At Nakaumi Lake, for example, the variability in N2O flux due to tidal
position outweighs diurnal and temporal variability [Hirota et al., 2007]. However, at field sites with high natural or anthropogenic nitrate concentrations, the water column is the primary source of nitrate for sedimentary denitrifiers, [e.g. Adams et al., 2012; Dong et al., 2006; Dong et al., 2002; Robinson et al., 1998; Tong et al., 2013] especially in highly bioturbated sediments [Law
et al., 1991]. Aerial exposure removes this nitrate source, slowing N2O production and emission, in particular in salt marshes and intertidal mudflats [Adams et al., 2012; Robinson et al., 1998].
As mentioned previously, N2O concentrations are typically greatest at upstream locations
within the estuary [Bange et al., 1996; Zhang et al., 2010]. However, in many estuaries the spatial pattern reflects a combination of environmental factors. For example, the Adyar River, in India, has a mid-estuarine N2O maximum [Rajkumar et al., 2008], and many European estuaries exhibit a spike in N2O emissions within a narrow salinity band, near the turbidity maximum
[Abril et al., 2000; Bange, 2006; Barnes and Owens, 1999; De Bie et al., 2002; De Wilde and De
Bie, 2000]. In these cases, N2O efflux responds to the residence time and movement of DIN through the catchment, as well as the concentration of suspended particulates. Other studies note the presence of N2O “hot spots.” [Corredor et al., 1999; Marty et al., 2001; Wong et al., 2013] or
find little correlation between N2O and salinity [Bange et al., 1998; Zhang et al., 2010]. In some cases, groundwater can be a confounding factor. DIN and N2O concentrations
are usually much higher in groundwater than in receiving water [Burnett and Dulaiova, 2003] and groundwater can be a significant N2O source into estuaries [LaMontagne et al., 2003; Wong
et al., 2013]. Tidal pumping of groundwater can also contribute to tidal variability in N2O
concentrations in the water column [Wong et al., 2013]. Groundwater N2O concentrations increase with increasing subterranean denitrification in nearby aquifers [LaMontagne et al.,
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2003; Smith et al., 1991], and buried sediments [Addy et al., 2013]. The potentially large effect of inundation and groundwater on N2O flux suggests a need for temporal and spatial sampling that includes the whole tidal cycle and more complete spatial coverage (near-shore, intertidal, subtidal etc.). It is also important to account for all of the physical emission pathways of N2O flux. N2O
can move through soil primarily through diffusion, but also through convection and advection [Clough et al., 2005]. However, static chamber and incubation N2O flux measurements only capture the N2O increase in a headspace, and do not differentiate between mechanisms of transport. In the discrete water sampling method, the calculated fluxes sum advection and diffusion of N2O from the water surface, but this does not consider ebullition [Clark et al., 1995]. In the case of CH4, ebullition contributes a significant portion of the water-air flux [Leifer and
Patro, 2002]. For N2O, on the other hand, ebullition may be a negligible effect. In freshwater
environments, for example, ebullition seems to contribute less than 0.1% of total N2O emissions
[Baulch et al., 2011]. However, considering that ebullition can cause erratic, non-linear changes in gas fluxes [Barnes et al., 2006; Rajkumar et al., 2008; Shalini et al., 2006], it is possible that
any chamber measurements affected by ebullition would be considered erroneous and discarded.
In addition to these mechanisms, dissolved N2O in pore water can be absorbed by plants
and emitted through transpiration [Chang et al., 1998]. In one study, plant transpiration was contributed up to to 43% of total N2O fluxes in salt marshes [Cheng et al., 2007]. It is common to place static chambers over plant life, so in most cases these flux measurements represent the plant plus sediment N2O fluxes. A few studies attempt to measure pneumatophore and plant stem emissions directly [Cheng et al., 2007; Hirota et al., 2007; Kreuzwieser et al., 2003; Krithika et al., 2008; Purvaja et al., 2004]. The effect of mangrove pneumatophores can be negative or positive, depending on the field location [Kreuzwieser et al., 2003; Krithika et al., 2008]. In one study static chambers covering mangrove pneumatophores exhibited fluxes 45% lower to 30% higher than chambers covering only bare sediment [Kreuzwieser et al., 2003]. In terrestrial and wetland environments, macrophyte arenchyma act as a conduit for the transfer of N2O and other
gases between the rhizosphere and the surface [Jorgensen et al., 2012; Yu and Chen, 2009]. In addition to transpiration, this is a likely contributor to the enhanced N2O emissions observed
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from salt marsh plants [Cheng et al., 2007] and from pneumatophores at some sites [Kreuzwieser et al., 2003; Krithika et al., 2008]. These macrophyte effects are complicated, however, by the competition between plants
and bacteria for DIN. This competition can result in lower soil-to-air N2O fluxes during the season of high macrophyte growth [Dausse et al., 2012; Yu et al., 2012] and in at least one case this was responsible for a negative correlation between DIN and N2O flux [Sun et al., 2013]. Without the competition from macrophytes, denitrifiers can produce more N2O, resulting in a
higher overall N2O flux, in some cases overshadowing plant transpiration. For example, clipping
salt marsh reeds reduced competition with sediment denitrifiers, resulting in a 30% to 100% increase in N2O fluxes in several studies [Cheng et al., 2007; Yu et al., 2012]. This research is
preliminary, and more work is needed to shed light on the complex plant-bacteria interactions, and their relation to N2O emissions. For example, it is understood that the C:N ratio of organic matter influences this plant-bacteria interaction. At high C:N ratios, mineralizing bacteria supplement organic N sources with inorganic N (ammonium and nitrate), limiting the inorganic N pool available to plants [Hodge et al., 2000]. This could limit denitrification, and potentially reduce N2O fluxes. In addition, the source of organic material also influences the amount of denitrification for a given amount of respiration [Eyre et al., 2013; Oakes et al., 2011] but it is unknown how organic matter quality influence N2O production. Significant diurnal variations in N2O fluxes are common. N2O emissions generally peak
at night, most likely due to increased NH4+, which can fuel nitrification and coupled nitrificationdenitrification [Bauza et al., 2002]. Dissolved ammonia is less abundant during the daylight hours because it is consumed by photosynthesizers [Biswas et al., 2005]. In addition, low night-
time dissolved oxygen enhances allows for denitrification closer to the sediment surface, enhancing N2O fluxes into the water column [Jensen et al., 1984]. These effects can be significant. Yu et al., [2012] found negative fluxes as low as -1.73 in Scirpus mariqueter plots during the summer season under light conditions, and positive fluxes with an average of 4 µmol m-2 h-1 under dark conditions. Similarly, in seagrasses, denitrification rates are higher under dark,
rather than light, conditions [Eyre et al., 2011].
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The seasonal variability of N2O fluxes can also be significant. N2O emissions are generally higher during the warmer season [Allen et al., 2011; Usui et al., 2001; Wang et al., 2007], most likely due to increased bacterial productivity at higher temperatures. Likewise, seagrasses in subtropical estuaries exhibit higher rates of denitrification [Eyre et al., 2013; Eyre et al., 2011; Ferguson et al., 2004] than in temperate estuaries [Risgaard-Petersen et al., 1998;
Welsh et al., 2001]. However, crop fertilization and plant growth cycles also play a role. For example, Dausse et al. [2012] found that N2O fluxes were negative during the season of plant matter decomposition, as a result of anoxic marsh waters and high NH4+ concentrations. Since
wind speed is a variable in the water-air N2O flux calculation, estuarine open water N2O fluxes
are often greater in the high-wind season [Gonçalves et al., 2010]. Additionally, precipitation washes sediment and DIN into the water column [Rajkumar et al., 2008] affects redox conditions [Ueda et al., 2000], and increases sediment moisture and density [Chauhan et al., 2008; Krithika
et al., 2008].
The spatial variability of N2O fluxes is high due to variations in DIN, oxygen, substrate
for bacteria, and other biogeochemical parameters [Gonçalves et al., 2010; C J Smith et al., 1983]. For example, landward sites seem to emit greater amounts of N2O than coastal and down-
river sites [Chen et al., 2012], possibly due to increased areal exposure, which leads to faster
N2O diffusion from the sediment to the atmosphere [Bauza et al., 2002]. However, sparse data limits our understanding of these patterns; spatial and temporal variability within these systems could be better characterized by more continuous and extensive monitoring.
Estimates of the global N2O flux from tidal environments The highest median fluxes of N2O were found in mangrove and estuarine open water
environments (Table 6), with lower rates from salt marsh and seagrass environments.
Interestingly, all N2O flux measurements are higher than the estimated area-weighted average of about 0.31 µmol m-2 h-1 for terrestrial soils [Chapuis-Lardy et al., 2007]. N2O fluxes vary
significantly; while the median N2O flux from all environments was about 0.59 µmol N2O m-2 h-1, median values between 3 and 7 µmol N2O m-2 h-1 are observed in systems with high DIN
concentrations [Barnes and Owens, 1999; Konnerup et al., 2014; Wong et al., 2013; Yu et al.,
2012], and individual measurements can reach 55 µmol N2O m-2 h-1 [Robinson et al., 1998].
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Considering the high variability and the sparse data coverage, any global estimate of estuarine N2O emissions will have a high uncertainty. Past estimates of the total open-water N2O
emissions involve multiplying dissolved N or total N concentrations by N2O emissions factors [Ivens et al., 2011; Kroeze et al., 2005; Kroeze et al., 2010; Seitzinger and Kroeze, 1998]. The N concentration data is often derived from complex models of catchment nutrient dynamics, and the emissions factors take into account observations of N2O:N2 ratios and bacterial N2O production rates. These estimates do not address N2O fluxes from vegetated tidal areas such as mangroves and salt marshes.
Another approach involves up-scaling direct N2O flux measurements from estuarine
environments. Several studies have estimated global mangrove sediment N2O emissions based on static chamber flux measurements from a few field sites [Chauhan et al., 2008; Corredor et
al., 1999; Robinson et al., 1998]. Essentially, these bottom-up estimates multiply local fluxes by estimates of the global mangrove areal extent. A similar approach has been used to estimate global N2O fluxes from the open water of estuaries, but with the inclusion of field data from more sites [Bange et al., 1996; Barnes and Upstill-Goddard, 2011; Law et al., 1992]. This method has the advantage of a greater reliance on field N2O flux measurements, however as of yet there are no such estimates from salt marsh, seagrass, or intertidal mudflat environments.
Here we present a simple up-scaling of reported values, in which local N2O emissions are
multiplied by the estimated global surface area of each environment type. For the purposes of this calculation, we use the median values of observed N2O flux ranges, where reported. If
median values are not available, N2O flux values are assumed to lie at the middle of the reported ranges. The midpoint of this data is then multiplied by the estimated areal extent of the respective environment type. In addition, approximate minimum and maximum N2O fluxes are identified from the 25% and 75% quartile values of the reported median fluxes (Table 6).
Mangrove forests occupy between 13.8 to 15.2 Mha of area worldwide [Giri et al., 2010;
McLeod et al., 2011; Pendleton et al., 2012; Spalding, 2010]. Using the median N2O flux of 0.95 µmol N2O m-2 h-1 (Table 1) and an area of 14.5 Mha [Pendleton et al., 2012] gives a global N2O flux of about 0.03 Tg N2O-N yr-1 from mangrove forests. The lower estimate of mangrove area
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(13.8 Mha) and the 25% quartile N2O flux estimate (0.4 µmol N2O m-1 h-1) give in a lower bound of 0.01 Tg N2O-N yr-1 emitted from mangrove forests. Similarly, the maximum areal extent (15.2 Mha) and 75% quartile value (2.4 µmol m-2 h-1) result in an upper estimate of 0.09 Tg N2O-N yr1
. However, recent findings that undisturbed mangroves may be sinks for N2O [Erler et al., in
press] suggest these estimates may be too high.
Little is known about the global areal extent of salt marshes, but a recent review by
Pendleton et al. [2012] approximates the global “tidal marsh” area at 5.1 Mha, and notes that a wide possible range of values [2.2 and 40 Mha] are reported [Chmura et al., 2003; Duarte et al., 2005; McLeod et al., 2011]. Using the “tidal marsh” area estimate, the global salt marsh N2O flux is calculated at 0.004 Tg yr-1, with a N2O flux range between 0.001 to 0.14 Tg N2O-N yr-1. We estimate the global open water estuarine area at 102.5 Mha, an average of the
estimates by Dürr et al. [2011] and Woodwell [1980]. Using the median water-air flux in estuarine waterways of about 0.76 µmol N2O m-2 h-1 gives an estimated global flux of 0.23 Tg N2O-N yr-1. The interquartile range of flux measurements (0.4 to 1.3 µmol N2O m-2 h-1) gives a
global flux range of 0.13 to 0.44 Tg N2O-N yr-1. No estimate currently exists of the global areal extent of estuarine intertidal sediments.
However, several studies have provided estimates for individual systems with a range from 10% to 20% of the open water estuarine area [Borges and Abril, 2011; Eyre et al., 2011; Zhao et al.,
2008]. Assuming that about half of mudflat area is submerged at any given time, we estimate a
global average exposed mudflat area of about 7 Mha, with a range from 5.5 to 14 Mha. Based on this range of values, we calculate global N2O fluxes between 0.01 and 0.05 Tg N2O yr-1. Seagrasses cover about 30 Mha of area, with a range from 17.7 to 60 Mha [Pendleton et
al., 2012]. There are no measurements of N2O fluxes from the sediments or water of seagrass
environments, but denitrification rates range from 0 to 800 µmol N m-2 h-1, with a median of 70
µmol N m-2 h-1 (Table 5). In benthic sediments, the ratio of N2O to N2 produced by denitrification ranges from about 0.1% to 6%, with the highest values found in environments with high DIN concentrations [Jensen et al., 1984; Nishio et al., 1983; Seitzinger, 1988]. There
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are no direct measurements of N2O/N2 ratio from denitrification in seagrasses, however considering that seagrasses thrive in less-disturbed estuaries [Short and Neckles, 1999], the ratio
is likely near the lower end of the benthic sediment range. Assuming a N2O/N2 ratio between 0.1% and 1%, we calculate a global N2O flux in seagrass of about 0.03 Tg N2O-N yr-1, with a
possible range between 0 and 0.19 Tg N2O-N yr-1. It should be noted that estimates of seagrass areal extent include seagrasses growing in marine continental shelves as well as estuaries, so this may be a slight over-estimate of the estuarine source.
The open water flux estimates fall close to an estimate by Robinson et al. [1998] of 0.13-
0.45 Tg N2O-N yr-1, as well as the value from Law et al. [1992] of 0.44 Tg N2O-N yr-1. In contrast, [Bange et al., 1996] report much higher values, between 3.7 and 5.7 Tg N2O-N yr-1, and
Kroeze et al. [2005] estimate N2O fluxes from combined river and estuarine open waters at about ~1.5 Tg N2O-N yr-1. The mangrove estimate, of 0.03 Tg N2O-N yr-1, is lower than a previous
estimate by Chauhan et al. [2008] of 0.11 Tg N2O-N yr-1 in mangroves in India, which received higher anthropogenic DIN loads. However Corredor et al. [1999] estimate a range of ~ 0.0140.56 Tg N yr-1, in a study of mangrove forests with both high and low concentrations of DIN.
The wide range of estimated global fluxes is likely due to the different methods employed.
The up-scaling estimates [e.g. Law et al., 1992; Robinson et al., 1998] tend to be conservative, while the emission-factor estimates [Kroeze et al., 2005] tend to report higher values. It may be the latter studies use DIN concentrations and emission factors are slightly too high. Alternatively, the up-scaling estimates are based on data with incomplete spatial and temporal coverage, so they may not take into account some of the peaks and “hot spots” of N2O production. The high estimates of Bange et al. [1996] may result from over-representation of a few
large, eutrophic European and U.S. estuary waters, including the Shelde and the Potomic. Bange et al. [1996] calculated global estuarine open water fluxes from aqueous N2O concentrations,
using average values rather than median values, as used here. In addition, most of the data was collected during the summer season, when N2O fluxes are generally higher. Other studies employing a similar approach found very different results [Butler et al., 1989; Elkins et al., 1978; Nevison et al., 1995] depending on the choice of air-sea exchange model.
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The combined N2O contribution of all estuarine environments is about 0.31 Tg N2O-N yr1
to the atmosphere, within an estimated range of 0.15 to 0.91 Tg N2O-N yr-1. It should be noted
that the total fluxes in general and the estuarine open water fluxes in particular may be biased slightly toward higher values, considering that many field sites were located in disturbed, highpopulation areas of Europe and Asia. As discussed before, the tendency toward warm-season and low tide sampling may also bias the global estimate.
At present, the total estuarine area contributes a relatively small positive flux of N2O. For
further comparison, the total natural and anthropogenic N2O source is estimated 17.3-17.7 Tg N
yr-1 [Denman et al., 2007], with an estimated open ocean source of 3.8– 5.8 Tg N yr-1 [Hirsch et al., 2006; Nevison et al., 2003]. The total source from agriculture is estimated at about 5.3 Tg N2O-N yr-1 [Alfi Syakila and Kroeze, 2011].
The future of estuarine N2O emissions The existing research indicates that the most important factors controlling N2O fluxes are
dissolved inorganic nitrogen (DIN) and oxygen availability, which in turn are affected by tidal cycles, groundwater, and macrophyte density. Global environmental change will modify these factors, which in turn should change the flux of N2O fluxes in estuarine environments (figure 4).
Nitrogen Over-Enrichment As discussed previously, N addition experiments and observational records consistently
document increased production and decreased consumption of N2O following N. This is consistent with the chemistry underlying denitrification; N2O is more likely to be used as an
electron receptor when other preferred receptors (NO3- and NO2-) are scarce [Betlach and Tiedje,
1981]. Projections of river and estuary N loading consistently indicate rising DIN concentrations globally [Seitzinger et al., 2010] which should result in a significant increase in N2O fluxes (figure 3). In sediments, the rate of this increase varies widely [Kreuzwieser et al., 2003; MuñozHincapié et al., 2002; Whigham et al., 2009]. However, in open water estuarine environments, it
seems that the N2O flux increases at a rate of 0.03 µmol m-2 h-1 N2O for every 1.0 µmol l-1 increase in nitrate concentration (figure 3a). A doubling of current median NO3- concentrations,
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from ~50 uM to ~100uM, would be expected to increase the global open water estuary N2O flux by about 0.45 Tg N2O-N yr-1, or 190%, for a total of about 0.68 Tg yr-1 (Figure 4). Nitrogen enrichment of estuaries also drives eutrophication [Cloern, 2001]. The
associated increase in phytoplankton production will increase the deposition of phyto-detritus to the sediments. Low C:N phyto-detritus may initially stimulate benthic denitrification and associated N2O fluxes [Eyre and Ferguson, 2009 ; Oakes et al., 2011]. However, high phytodetritus loadings will reduce N loss via denitrification and enhance recycling of N to the water column as NH4 [Eyre and Ferguson, 2009] . This in turn may decrease N2O fluxes, but it is not
entirely clear what will be the overall effect on N2O fluxes. It is unlikely that N loading will increase uniformly in every estuary system. River and
estuary nutrient and N2O concentrations will likely increase quickly in rapidly developing
countries, such as China and India, due to population increase and fertilizer use [Nixon, 2009]. In contrast, some European or U.S. estuaries may see a decrease in DIN concentrations, owing to slow population growth, upgraded sewage treatment and other environmental mitigation efforts [Mozetic et al., 2010; Stockner et al., 2000]. Mitigation may lead to a decrease in N2O emissions,
or may move “hot spots” of N2O production and consumption further inland or further upstream [Burgin et al., 2013]. Additional data is needed to fully account for these differences in N loading through time. In particular, projections of future DIN concentration need to take into account the areal extent of estuarine environments – open water, intertidal, etc – in each country affected by different N loading scenarios.
Hypoxia Increasing DIN loads will likely cause a spread of hypoxic and anoxic zones in estuaries
[Howarth et al., 2011]. Additionally, loss of vegetation and loss of habitat for benthic burrowing macrofauna may lead to decreased delivery of O2 by roots and bioturbation, favoring anaerobic
N2O production by denitrification. While decreased O2 can increase the bacterial consumption of N2O by denitrifiers, under high NO3- concentrations the N2O:N2 ratio of denitrification remains
high [Muñoz-Hincapié et al., 2002; Seitzinger and Nixon, 1985; Teixeira et al., 2010] suggesting that production of N2O significantly out-paces consumption. Where NH4+ dominates the DIN
pool, hypoxia may lead to negative N2O fluxes, as seen in some hypoxic bottom waters
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[LaMontagne et al., 2003; Rajkumar et al., 2008], however it is likely that hypoxia combined with abundant NO3- will increase in N2O emissions. It is difficult to estimate the magnitude of the O2 effect on overall N2O emissions, as it
seems to be non-linear and there are interactions with DIN concentrations. However, it will most likely be a positive effect. In the ocean, low-oxygen seawater is associated with higher N2O
concentrations [Naqvi et al., 2010; Zhang et al., 2008]. Water-column N2O production is highest at intermediate O2 saturations, between 2 and 15% [De Bie et al., 2002]. In a controlled
experiment, O2 saturation of about 2% was observed to increase steady-state N2O concentrations in a headspace to a peak of 12ppm from a background concentration less than 2 ppm [De Bie et al., 2002].
Decreased macrophyte distribution Global change is rapidly altering the type, and function, of primary producers in coastal
ecosystems with a reduction in mangroves, seagrasses and salt-marshes by around 30 to 50% over the past few decades due to reclamation, deforestation and urbanisation [McLeod et al., 2011]. Macrophytes can act as a conduit of N2O from the sediment to the atmosphere, and hence removing vegetation may decrease N2O emissions in some areas. However, clearing of coastal
vegetation will also lead to a higher concentration of DIN in estuarine waters and associated increased N2O fluxes (see above), and will decrease plant competition for DIN, potentially allowing for greater bacterial production of N2O. Change to organic matter quality associated
with different primary producers influences the amount of benthic denitrification for a given amount of respiration [Eyre et al., 2013]. However, it is unknown how these changes in organic matter quality influence N2O production. Although attempts have been made to mitigate enhanced N loads through rehabilitation
of vegetated areas such as “managed realignment” areas [Andrews et al., 2006; French, 2006], unexpectedly high N2O emissions have been measured in managed realigned salt marshes [Adams et al., 2012; Blackwell et al., 2010] suggesting these environments may mitigate high N loadings at the expense of significant N2O emissions.
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When compared with adjacent subtidal sediments, denitrification rates were found to be between 6 and 7 times higher in seagrasses [Eyre et al., 2013; Piehler and Smythe, 2012]. Assuming this relationship is typical of estuarine areas, this would imply a significant decrease in benthic N2O production after seagrass loss. If 50% of seagrass cover were lost, an estimated net decrease of 0.012 Tg N2O-N yr-1 would result (Figure 4). A 100% loss of seagrass would
result in an estimated decrease in N2O production of about 0.024 Tg N2O-N yr-1. However, this
is assuming that the N2O:N2 ratios of denitrification in seagrasses are typical of those found in other benthic sediments. It is possible that the N2O:N2 ratios are low in seagrasses, considering
that seagrasses thrive in pristine, low-DIN estuaries. Additional data is needed to further refine this estimate.
A loss of mangrove vegetation could result in an increase in unvegetated intertidal area,
resulting in a slight decrease in N2O emissions in these environments. A 50% loss of mangrove vegetation cover would result in a decrease in N2O emissions of 0.017 Tg N2O-N yr-1. If that area were converted to unvegetated intertidal bare sediment this would increase the N2O flux
from these unvegetated environments by 0.015 Tg N2O-N yr-1, for a net decrease of 0.002 Tg
N2O-N yr-1. It is not clear that intertidal areas will necessarily be converted to unvegetated land cover.
Coastal development and land reclamation could lead to a loss of vegetated tidal areas as they are buried under fill material or converted to grazing pastures. It is difficult to know what effect this
will have on N2O emissions. Denitrification and N2O production still occur under buried salt marsh at and below the buried sediment horizon [Addy et al., 2013]. Thus, N2O produced below
ground may still be delivered to adjacent areas and emitted. In estuarine groundwater with lownitrate concentrations and high pH, N2O:N2 ratios are very low and N2O contributions are likely
small [Addy et al., 2005]. However, N2O production and N2O;N2 ratios in groundwater will likely increase in the future, as increasing NO3- concentrations promote denitrification and
inhibit N2O consumption [Smith et al., 1991; Addy et al., 2013].
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Grazing may have a minimal impact on overall marsh N2O emissions [Ford et al., 2012], with the primary effect being N mineralization rates by trampling sediment and removing litter [van Wijnen et al., 1999]. However, in these studies, the herbivores were small wild species (e.g. hares, geese). Domestic ruminants are known to increase N2O emissions significantly through excretion of high N waste onto the soil surface. N2O emission rates from pasture land in New
Zealand range from 1.6 µmol m-2 h-1 from sheep grazing to a maximum of 4 µmol m-2 h-1 from cattle grazing [Saggar et al., 2008], much higher than our estimated salt marsh average of 0.34 µmol m-2 h-1. If 50% of salt marsh area were lost, this would decrease salt marsh N2O emissions
to about 0.0022 Tg N2O-N yr-1 (Figure 4). However, an equivalent area of pasture would produce between 0.01 and 0.025 Tg N2O-N yr-1, up to 10 times this amount. The same area of
unvegetated intertidal flats would produce only about 0.003 Tg N2O-N yr-1, for a net increase of
0.001 Tg N2O-N yr-1. If intertidal unvegetated areas, such as mudflats, replaced 50% of both mangrove and salt marsh, the overall N2O flux from mudflats would increase to 0.024 Tg N2O-N yr-1 (Figure 4).
Climate Increased temperature associated with climate change generally increases the rate of
bacterial activity and gaseous diffusion, and may lead to increased N2O production and emission
from soils [Barnes and Owens, 1999; Nowicki, 1994]. However, the effect of temperature on the
yield of N2O vs. N2 from denitrification in estuarine environments remains unclear. It is possible that other climatic changes will have a greater effect, for example changes in evaporationprecipitation balance will likely to affect soil moisture, and changes in wind speed will affect the N2O flux from open waters. Additional research is needed to better anticipate climate effects on
N2O emissions from estuaries. Conclusions Estuarine environments are currently a small source of N2O, 0.31 Tg N2O-N yr-1 or less
than 2 % of the total anthropogenic source [Denman et al., 2007]. At present, incomplete spatial and temporal sampling limits our understanding of the biogeochemical factors that affect N2O
flux rates. In particular, the seasonal and diurnal patterns of N2O emission seem to differ from one site to another. Nonetheless, several factors, such as DIN concentration, O2 concentration,
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macrophyte density, salinity, and groundwater, strongly and consistently affect the sediment-toair and water-to-air N2O fluxes. Future global changes in climate and anthropogenic N loading are expected to increase
N2O emissions from estuaries. Production of N2O by nitrification may be inhibited slightly by
decreased O2 and changes in macrophyte cover, but denitrification activity is expected to increase, as is the ratio of N2O:N2 produced by denitrifiers. More research is needed to further
elucidate these relationships.
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Smith, K. A., I. P. McTaggart, and H. Tsuruta (1997), Emissions of N2O and NO associated with nitrogen fertilization in intensive agriculture, and the potential for mitigation, Soil Use and Management, 13(4), 296-304. Smith, R. L., B. L. Howes, and J. H. Duff (1991), Denitrification in nitrate-contaminated groundwater—occurrence in steep vertical geochemical gradients, Geochimica Et Cosmochimica Acta, 55(7), 1815-1825. Spalding, M. (2010), World Atlas of Mangroves, 107-109 pp., Earthscan, London, Washington D.C. Stocker, T. F., D. Qin, G.-K. Plattner, M. Tignor, S. K. Allen, J. Boschung, A. Nauels, Y. Xia, V. Bex, and P. M. Midgley (2013), Climate Change 2013. The Physical Science Basis. Working Group I Contribution to the Fifth Assessment Report of the Intergovernmental Panel on Climate Change-Abstract for decision-makersRep., Groupe d'experts intergouvernemental sur l'evolution du climat/Intergovernmental Panel on Climate Change-IPCC, C/O World Meteorological Organization, 7bis Avenue de la Paix, CP 2300 CH-1211 Geneva 2 (Switzerland). Stockner, J. G., E. Rydin, and P. Hyenstrand (2000), Cultural oligotrophication: Causes and consequences for fisheries resources, Fisheries, 25(5), 7-14.
Sun, Z. G., L. L. Wang, X. J. Mou, H. H. Jiang, and W. L. Sun (2014), Spatial and temporal variations of nitrous oxide flux between coastal marsh and the atmosphere in the Yellow River estuary of China, Environmental Science and Pollution Research, 21(1), 419-433. Sun, Z. G., L. Wang, H. Tian, H. Jiang, X. Mou, and W. Sun (2013), Fluxes of nitrous oxide and methane in different coastal Suaeda salsa marshes of the Yellow River estuary, China, Chemosphere, 90(2), 856-865. Svensson, J. M., A. Enrich-Prast, and L. Leonardson (2001), Nitrification and denitrification in a eutrophic lake sediment bioturbated by oligochaetes, Aquatic Microbial Ecology, 23(2), 177-186.
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Accepted Article
Syakila, A., and C. Kroeze (2011), The global nitrous oxide budget revisited, Greenhouse Gas Measurement and Management, 1(1), 17-26.
Syakila, A., C. Kroeze, and C. P. Slomp (2010), Neglecting sinks for N2O at the earth's surface: does it matter?, Journal of Integrative Environmental Sciences, 7, 79-87. Teixeira, C., C. Magalhães, R. a. R. Boaventura, and A. a. Bordalo (2010), Potential rates and environmental controls of denitrification and nitrous oxide production in a temperate urbanized estuary, Marine environmental research, 70(5), 336-342.
Tong, C., J. F. Huang, Z. Q. Hu, and Y. F. Jin (2013), Diurnal variations of carbon dioxide, methane, and nitrous oxide vertical fluxes in a subtropical estuarine marsh on neap and spring tide days, Estuaries and Coasts, 36(3), 633-642. Ueda, S., C. S. U. Go, T. Yoshioka, N. Yoshida, E. Wada, T. Miyajima, A. Sugimoto, N. Boontanon, P. Vijarnsorn, and S. Boonprakub (2000), Dynamics of dissolved O2, CO2, CH4, and N2O in a tropical coastal swamp in southern Thailand, Biogeochemistry, 49(3), 191-215. Usui, T., I. Koike, and N. Ogura (2001), N2O production, nitrification and denitrification in an estuarine sediment, Estuarine, Coastal and Shelf Science, 52, 769-781. Vachon, D., Y. T. Prairie, and J. J. Cole (2010), The relationship between near-surface turbulence and gas transfer velocity in freshwater systems and its implications for floating chamber measurements of gas exchange, Limnology and Oceanography, 55(4), 1723. Van Groenigen, J. W., K. B. Zwart, D. Harris, and C. van Kessel (2005), Vertical gradients of δ 15 N and δ18O in soil atmospheric N2O-temporal dynamics in a sandy soil, Rapid Communications in Mass Spectrometry, 19(10), 1289-1295. van Wijnen, H. J., R. van der Wal, and J. P. Bakker (1999), The impact of herbivores on nitrogen mineralization rate: consequences for salt-marsh succession, Oecologia, 118(2), 225-231. Volkenborn, N., L. Polerecky, S. Hedtkamp, J. Van Beusekom, and D. De Beer (2007), Bioturbation and bioirrigation extend the open exchange regions in permeable sediments, Limnology and Oceanography, 52(5), 1898-1909. Wang, D., Z. Chen, J. Wang, S. Xu, H. Yang, H. Chen, L. Yang, and L. Hu (2007), Summertime denitrification and nitrous oxide exchange in the intertidal zone of the Yangtze Estuary, Estuarine, Coastal and Shelf Science, 73(1-2), 43-53. Webb, A. P., and B. D. Eyre (2004), Effect of natural populations of burrowing thalassinidean shrimp on sediment irrigation, benthic metabolism, nutrient fluxes and denitrification, Marine Ecology Progress Series, 268, 205-220. Weiss, R., and B. Price (1980), Nitrous oxide solubility in water and seawater, Marine Chemistry, 8(4), 347-359.
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Accepted Article
Welsh, D., M. Bartoli, D. Nizzoli, G. Castaldelli, S. A. Riou, and P. Viaroli (2000), Denitrification, nitrogen fixation, community primary productivity and inorganic-N and oxygen fluxes in an intertidal Zostera noltii meadow, Marine Ecology Progress Series, 208(5).
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Yan, W., L. Yang, F. Wang, J. Wang, and P. Ma (2012), Riverine N2O concentrations, exports to estuary and emissions to atmosphere from the Changjiang River in response to increasing nitrogen loads, Global Biogeochemical Cycles, 26(4). Ye, Y., C. Lu, and P. Lin (2000), Seasonal and spatial changes of methane emissions from mangrove wetlands in Hainan and Xiamen, Chinese Journal of Atmospheric Science, 24, 152156. Yu, K., and G. Chen (2009), Nitrous oxide emissions from terrestrial plants: observations, mechanisms and implications, in Nitrous Oxide Emissions Research Progress, edited by B. E. Sheldon AI, pp. 85-104, Nova Science Publishers, Inc, New York, USA. Yu, Z., Y. Li, H. Deng, D. Wang, Z. Chen, and S. Xu (2012), Effect of Scirpus mariqueter on nitrous oxide emissions from a subtropical monsoon estuarine wetland, Journal of Geophysical Research: Biogeosciences, 117(G2), 1-9.
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Accepted Article
Zhang, G., J. Zhang, S. M. Liu, J. L. Ren, and Y. C. Zhao (2010), Nitrous oxide in the Changjiang (Yangtze River) Estuary and its adjacent marine area: Riverine input, sediment release and atmospheric fluxes, Biogeosciences, 7(11), 3505-3516. Zhao, B., H. Guo, Y. Yan, Q. Wang, and B. Li (2008), A simple waterline approach for tidelands using multi-temporal satellite images: a case study in the Yangtze Delta, Estuarine, Coastal and Shelf Science, 77(1), 134-142. Figure captions Figure 1. Locations of N2O flux measurements worldwide (a), in Europe (b), and in East Asia (c).
Figure 2. Latitudinal distribution of mangrove N2O flux field sites (from this review), and
mangrove areal extent (from Giri et al., 2001). Figure 3. Aqueous NO3- (a) and NH4+ (b) concentrations vs. water-to-air estuarine N2O fluxes. In the Humber River and Adyar River, average N2O fluxes are lower than expected for the given
DIN concentrations. Figure 4. Fluxes of N2O (Tg N2O-N yr-1) under current conditions (a), and under a future scenario (b). The future scenario considers N2O increase due to a doubling of DIN in main
estuary waters (but does not consider DIN changes in other environments) and the change in N2O fluxes due to 50% loss of mangroves, salt marsh, and seagrass.
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ccepted Articl
Mangroves Location
Method
Temporal coverage
Spatial coverage
N2O Flux range
Mean
Reference
-2
µmol N2O m h-1 Temperate Mornington Penninsula, Victoria, Australia
Static chamber
Four seasons sampled, with transects at the outgoing, incoming, and low tide
Three transects at three sites, with nine locations each, with Three environment types sampled
0.09
-
Two sites: near the water line and in the forest.
0.09
-
1.5
0.00
Four sites chosen along the Brisbane River and near the shore of Moreton Bay
0.15
-
4.6
0.00
Cores collected in summer and fall
Three Stations near the mouth of the estuary
0.23
-
0.48
0.00
0.33
Alongi et al., 2005
Measurements taken in winter and spring
Five sites
0.45
-
0.32
0.00
0.09
Kreuzweiser et al., 2003
Mangrove forest and adjacent creek
0.12
-
1.3
0.00
Barnes et al., 2006
Three sites on a transect normal to the shore
0.5
-
1.4
0.00
Bauza et al., 2002
0.9
-
4.7
1.5
0.44
Livesley and Andrusiak, 2012
0.44
Allen et al., 2007
Subtropical Moreton Bay, QLD, Australia
Static chamber
Moreton Bay, QLD, Australia
Static chamber
Jiulongjiang Mangrove, Xiamen, China
Lab incubation with air sampling Static chamber, lab incubation with air sampling
Moreton Bay, Woody Isle, and Port Douglas, QLD, Australia
Spring, winter, and summer 24hour series, sampled at three-hour intervals Two four-day periods in summer in winter, samples taken during low tide
Allen et al., 2011
Tropical Wright Myo, Andaman Island
Static chamber, floating chamber, discrete water samples
Static chambers deployed at low tide. Floating chamber deployed at high tide Full tidal cycle sampled at threehour intervals
Magueyes island, Southwest Puerto Rico
Static chamber
Bhitarkanika, India
Static chamber
Pre-monsoon, monsoon, postmonsoon seasons sampled
12 sites in the forest and near the main channel
Mangroves near Shenzhen and Hong Kong
Static chamber
Sampling performed two hours before low tide
Four sites sampled at three tidal positions, with three chambers each
0.00
-
0.1
0.00
0.1
Sai Keng Mangrove, Hong Kong
Lab incubation with air sampling
sampling for several weeks
Sediments collected at one site
0.23
-
0.48
0.00
0.33
Mai Po Mangrove, Hong Kong
Static chamber
Sampling over four seasons two hours before low tide
Four tidal positions sampled from the dense forest to the fringe mudflat
0.73
-
12.1
0.00
4.4
Chen et al., 2012
Southwest Puerto Rico
Static chamber
Four field sites, with one site sampled in a transect normal to the shore
0.12
-
7.8
0.00
1.7
Corredor et al., 1999
The Caribbean coast of Colombia
Static chamber
Seasonal coverage varied, samples collected in the morning Sampling performed monthly between 10:00 and 16:00 h over 11 months
0.8
-
26.8
8.7
Konnerup et al., 2014
Muthupet Mangrove, India
Static chamber, discrete water samples
Sampling over five seasons
15 sites along the length of the river
0.41
-
0.77
0.6
Krithika et al., 2008
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Four study sites selected based on conservation status of mangroves
Chauhan et al., 2008
0.00
Chen et al., 2010 Chen et al., 2011
ccepted Articl
Southwest Puerto Rico
Static chamber
Measurements taken at at high tide and low tide
One site near low tide mark and One site further upland
0.02
-
Pichavaram Mangrove, India
Unknown
Unknown
Unknown
0.89
-
1.8
North Hutchinson Island, Florida, USA
Static chamber
One measurement made per plot every two weeks over three years
12 study plots
0
-
0.46
0.1
0.00
0.5
Muñoz-Hincapié et al., 2002 Senthilkumar et al., 2008
0.11
Whigham et al., 2009
Table 1. Summary of mangrove N2O flux methods and results. Shaded studies appear in more than one environmental category.
Salt Marsh Location
Method
Temporal coverage
Spatial coverage
N2O Flux range
Mean N2O Flux
Reference
µmol N2O m-2 h-1 Temperate
Blackwater Estuary, UK
Static chamber
Four seasons sampled
River Torridge, Devon, UK
Lab incubation with air sampling
Three tidal cycles simulated
Tubul and Raqui Estuary system, central Chile
Discrete water samples
River Dovey, Wales, UK
Static chamber
Rio San Pedro, Gulf of Cadiz, Spain
Discrete water samples
Ribble Estuary, Northwest England
Static chamber
Lake Nakaumi salt marsh, Japan
Static chamber
Couesnon River, France and Torridge River, UK
Lab incubation with air sampling
Mornington Penninsula, Victoria, Australia
Static chamber
Rowley, Massachusetts
Static chamber
Four natural and managed realignment sites Two natural and managed realigment sites
-1.5
-
5
0.52
-3.2
-
21
Seven sites along the Tubul River, and four along the Racqui River
2.46
-
1.17
Four sites in fresh, brackish and salt marsh
-20
-
5
1.0
-
2.6
Adams et al., 2012 Blackwell et al., 2010
Summer and winter seasons sampled, as well as one 12-hour timeseries Spring and fall seasons sampled. Measurements taken at low tide Ten 24-hour sampling campaigns performed from winter to fall seasons Sites were sampled monthly for one year, during daylight hours
Six grazed and Six ungrazed plots along a fence line
0.07
-
1.1
0.08
Ford et al., 2012
Gas samples taken every six hours over two days in summer
Salt marsh and sandy shore sampled in a transect normal to the water line
0.68
-
1.4
~0.03
Hirota et al., 2007
0.1
-
0.7
0.5
Kenny et al., 2004
0
-
0.4
0.2
Livesley and Andrusiak, 2012
-6.1
-
0.02
-1.3
Moseman-Valtierra et al., 2011
Sediment sampled during summer, tidal cycles simulated Four seasons sampling, at the outgoing tide, at low-tide on the incoming tide Sampling at low tide, mid-day at spring tide in spring and summer seasons
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One fixed location
Two estuaries sampled Three transects at three sites, with nine locations each in three environment types Five sample sites in a line across the marsh flat.
0.00
-0.10
Daniel et al., 2013
-2.5
Dausse et al., 2012 Ferrón et al., 2007
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Yellow River Delta, China
Yellow River Delta, China
Static chamber, with plants included in the marsh chambers Static chamber, with plants included in the marsh chambers
Four seasons sampled during daylight hours at low tide
Three different tidal positions sampled: high, low and middle marsh
0.33
-
2.2
0.39
Sun et al., 2013
Four seasons sampled during daylight hours at low tide
Three different tidal positions sampled: high, low and middle marsh
0.11
-
1.8
0.30
Sun et al., 2014
-5.8
-
6.6
2.2
Wang et al., 2007
7.0
Yu et al., 2012
Yangtze (Changjiang) River, China
Static chamber
Measurements taken every 1.5 hours during the emergence (low tide) time of day
One site at two locations: high marsh and bare mudflat
Yangtze (Changjiang) River, China
Static chamber
Samples collected from May to October during daylight hours
One site in Scirpus mariqueter zone
Measurements taken about every six weeks over two years.
Three sites in fresh, brackish, and salt marshes
0.01
-
1.0
0.18
Smith et al., 1983
24-hour series sampled during spring and neap tides in spring and autumn seasons
Floating chambers deployed in the water near the static chambers
-5.1
-
12.6
1.4
Tong et al., 2013
Gulf Coast, USA
Subtropical Static chamber, lab incubation with air sampling
Min River Estuary, Southeast China
Static chamber, floating chamber
Table 2. Summary of salt marsh N2O flux methods and results. Shaded studies appear in more than one environmental category.
Estuarine open water Location
Method
Temporal coverage
Spatial coverage
N2O Flux range
Mean Flux
Reference
-2
µmol N2O m h-1 Temperate The Bodden waters near the Oden Estuary, Germany
Continuous water sampling
Four seasons sampled in five campaigns over several years
19 sites in four sub-basins within the Bodden waters
Discrete water samples
Spring and summer seasons sampled
Five sites along the estuary
Discrete water samples
Cruise undertaken in spring
Seven estuaries sampled along a transect the mouth to the limit of tidal influence
0.10
-
3.2
Alsea Estuary, Oregon, USA
Discrete water samples
Five sampling campaigns in summer and autumn seasons
Six sites along the length of the estuary
0.05
-
0.72
0.23
De Angelis and Gordon, 1985
Shelde River
Discrete water samples, continuous water samples
12 "compartments" sampled along the length of the estuary
~0.4
-
22
8.26
De Wilde and De Bie, 2000
10 estuaries in the UK
Discrete water samples
10-16 sites per river
0.18
-
11.1
Humber Estuary and Tweed Estuary, UK Humber, Forth, Tamar, Tyne, Tees, Wash, and Tay Estuaries, UK
Surveys performed in October, July, and March Water samples taken at high tide over four seasons
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0.0144
-
0.30
0.137
Bange et al., 1998
1.80
Barnes and Owens, 1999 Barnes and UpstillGoddard, 2011
Dong et al., 2004
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Twice monthly at inlet and outlet, and two to three times each summer for 11 years Three seasons sampled, 24-hour series measured at both spring and neap tides
Seine River Estuary, France
Discrete water samples
Tagus Estuary, Portugal
Discrete water samples
Tokyo Bay, Japan
Discrete water samples
May to October, 1994
Limfjorden, Denmark
Lab incubation with air sampling
Child's River, Massachusetts, USA
3.20
-
4.6
4.02
Garnier et al., 2006
-0.04
-
0.5
0.12
Goncalves et al., 2010
23 sampling sites sampled on a cruise of the bay
0.06
-
6.4
1.20
Hashimoto et al., 1999
Cores collected over four seasons, incubated for three to four hours
One site
0.17
-
1.8
0.95
Jensen et al., 1984
Discrete water samples
Summer and fall seasons sampled
Sampling along the length of the estuary
0
1.6
0.46
LaMontagne et al., 2003
Tamar River, UK
Discrete water samples
Four seasons sampled
Four transects along the estuary and one 24-hour tidal cycle sampled
0
-
1.5
0.40
Law et al., 1992
Po River, Italy
Discrete water samples
-0.18
-
6.6
0.90
Leip, 2000
Brisbane River, QLD, Australia
Discrete water samples
0.09
-
3.2
0.63
Musenze et al., 2014
River Colne, UK
Discrete water samples
0
-
55.8
1.30
Robinson et al., 1998
River Tyne, UK
Discrete water samples
Summer season sampled
14 sites along the river
Narragansett Bay, Rhode Island, USA
Lab incubation with air sampling
Cores collected between March and November
Three sites: offshore, mid-bay, and river mouth
Temmesjoki Fjord
Discrete water samples, floating chamber
Werribee Estuary, Victoria, Australia
Discrete water samples
Yangtze (Changjiang) River, China
Discrete water samples
Jiaozhou Bay, China
Discrete water samples
Four seasons sampled about once per month Monthly surveys performed for one year. Two 30-hour times series conducted 24-hour sampling study conducted in August and October Sampling in May, August, and December
Gulf Coast, USA
Adyar River
Subtropical Static chamber, lab incubation with air sampling Tropical Discrete water samples, floating chamber
Eight sampling campaigns over one year, low and high tide sampled Four seasons sampled over several years. A 24-hour series was also sampled Surveys performed monthly for one year, six surveys each at high and low tide
12 sampling sites along the estuary 15 water sampling stations along length of the river.
Six sampling sites 18 stations along the estuary sampled starting at high tide at the mouth of the estuary 20 water sampling sites along the length of the estuary
Rodrigues et al., 2007 0.02
-
0.1
Eight sites along the river and estuary Six sites along the length of the river and estuary
2.75
-
4.0
Two sites
0.07
-
1.5
Many sites sampled during five cruise campaigns
Measurements taken about every six weeks over two years.
Three sites in fresh, brackish, and salt marshes
Samples collected every two to three months over 1.5 years.
20 sites along the river and estuary
0.00
-
0.30
Table 3. Summary of estuarine open water N2O flux methods and results. Shaded studies appear in more than one environmental category.
This article is protected by copyright. All rights reserved.
Seitzinger et al., 1984 0.63
Silvennoinen et al., 2008
3.40
Wong et al., 2013 Yan et al., 2012
0.5
Zhang et al., 2006
0.1
Smith et al., 1983
0.9
Rajkumar et al., 2008
ccepted Article
Intertidal sediment Location
Method
Temporal coverage
Spatial coverage
Mean N2O Flux µmol N2O m-2 h-1
N2O Flux range
Reference
Temperate Four natural and managed realignment sites sampled
-1.2
-
2.7
0.65
Adams et al., 2012
Five sites along the length of the estuary
0.17
-
25.2
3.3
Barnes and Owens, 1999
0.1
-
0.4
0.2
Kenny et al., 2004
Six sampling sites
0
-
4.5
Sediment cores taken at low tide, four times over one year. Both dark and light conditions tested
Two sites - one in sandy sediment and one in muddy sediment
0
-
1.3
Static chamber
Sampling every four weeks over a 14-month period at low tide
Two plots each at nine sites along the estuary
-1.4
-
4.3
River Colne, UK
Static chamber, lab incubation with air sampling, discrete water samples
Incubations simulated tidal conditions, static chambers were deployed at low tide
Six cores taken from four sites along the estuary. Static chamber placed at low tide mark.
0.07
-
9.1
0.72
Robinson et al., 1998
Yellow River Delta, China
Static chamber
Four seasons sampled during daylight hours at low tide
One bare mudflat site sampled
0.26
-
1.2
0.36
Sun et al., 2013
Yellow River Delta, China
Static chamber
Four seasons sampled during daylight hours at low tide
One bare mudflat site sampled
0.01
-
0.9
0.31
Sun et al., 2014
Yangtze (Changjiang) River, China
Static chamber
Measurements taken every 1.5 h during low tide.
One site at two locations: high marsh and bare mudflat
-2.9
-
0.25
Blackwater Estuary, UK
Static chamber
Four seasons sampled
Humber Estuary and Tweed Estuary, UK Couesnon River, Normandy, France and Torridge River, Devon, UK
Lab incubation with water sampling
Po River, Italy
Static chamber
Cores collected during spring and summer seasons Cores taken during summer, tidal cycles simulated and sampled every 6 hours Eight sampling campaigns over one year
Douro River Estuary, Portugal
Lab incubation with water sampling
Sheldt River Estuary, Netherlands
Lab incubation with air sampling
Two estuaries sampled
Table 4. Summary of intertidal sediment N2O flux methods and results. Shaded studies appear in more than one environmental category.
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0.0
Leip, 2000 0.07
Magalhäes et al., 2005 Middelburg et al., 1995
Wang et al., 2007
ccepted Articl
Seagrass (denitrification measurements) Location
Method
Temporal coverage
Spatial coverage
Denitrification Mean range -2 -1 µmol N m h
Reference
Temperate Løgsør Bredning, Limfjorden, Denmark
Isotope pairing analysis of in-situ chamber incubations
Measurements were made in April and August, 24-hour incubations
One site in the eastern part of the
3
-
13
Risgaard-Petersen et al., 1998
Risgarde Bredning and Bight of Aarhus, Limfjorden, Denmark
Isotope pairing analysis of laboratory incubations
Measurements performed monthly over a one-year period Samples taken in September and January, incubated over for up to eight hours Samples taken in February, May and October, and incubated for one to two hours
Two sites: one sheltered and one windexposed
2
-
35
Risgaard-Petersen and Ottesen, 2000
Bassin d’Arcachon and Etang du Prévost, France
Isotope pairing
>1
-
12
Rysgaard et al., 1996
Ile aux Oiseaux, Bassin d’Arcachon, France
Isotope pairing analysis of laboratory incubations
One sampling station in an intertidal Zostera noltii bed
2
-
6
Welsh et al., 2000
Ile aux Oiseaux, Bassin d'Arcachon, France
Isotope pairing, Acetelyne block
Cores collected in October, incubated for one hour
One sampling station in an intertidal Zostera noltii bed
0
-
50
4
Welsh et al., 2001
Two estuary systems sampled
Subtropical Moreton Bay, Australia
N2: Ar analysis conducted ex-situ at the study site
Measurements were made in summer and winter with 24-hour incubations starting at sunset
One site each in Halophila and Zostera seagrass beds
28
-
824
360
Eyre et al., 2011
Wallis Lake, Camden Haven, and Hastings River estuary, Australia
N2: Ar analysis of in-situ and laboratory incubations
Cores incubated over a 24-hour diel cycle, four seasons sampled
Three estuaries sampled in Zostera, Halophila, Ruppia, and Posidonia seagrass beds
12
-
455
123
Eyre et al., 2013
Brunswick Estuary, Australia
N2: Ar analysis of on-site sediment core incubations
Six incubations performed over a 1.5-year time period
Four sites along the length of the estuary
524
Ferguson, Gay and Eyre, 2004
Tropical Oyster Bay, Falmouth Harbour, Jamaica
Acetelyne block method performed on laboratory sediment samples
Florida Bay, USA
N2: Ar
Samples collected in August and December and incubated for several hours Samples collected in August and March under both light and dark conditions
Table 5. Summary of seagrass denitrification methods and results.
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One sample site
six basins sampled within the bay
83
-
167
Blackburn et al., 1994
0
-
260
Kemp and Cornwell, 2001
ccepted Articl
Environment
Aerial extent
N2O flux measurements -2
(Mha)
(µmol N2O m h ) Inter-quartile Min range
Range
Mean
Median
Mangrove
13.8 to 15.2a
14.5b
0.95
0.4 to 2.4
Salt marsh
2.2 to 40c
5.1b
0.34
110 to 140d
125
Intertidal sediment
5.5 to 14e
Seagrassg Total
Estuarine open water
Global N2O flux estimates
-1
(Tg N2O-N yr-1) Inter-quartile Min range
Max
Median
0.1
6.0
0.03
0.01 to 0.09
0.002
0.22
0.1 to 1.5
-2.5
8.9
0.004
0.001 to 0.14
-0.01
0.88
0.76
0.4 to 1.3
0.1
8.3
0.23
0.13 to 0.44
0.01
2.8
9.8
0.50
0.3 to 1.4
0.1
3.3
0.012
0.01 to 0.05
0.001
0.11
17.7 to 60f
30
0.39
0.007 to 1.3
0.0
5.2
0.029
0 to 0.19
0.0
0.77
149 to 269
184
0.59
0.2 to 1.6
-2.5
8.9
0.31
0.15 to 0.91
0.005
4.8
Table 6. Median N2O fluxes and Global N2O flux estimates from estuarine environments. a. From Giri et al., 2010, McLeod et al., 2011, and Spalding, 2010 b. From Pendleton et al., 2012 c. From Chmura et al., 2003, Duarte et al., 2005, and McLeod et al., 2011 d. From Dürr et al., 2011 and Woodwell, 1980 e. Estimated as 1.6% to 20% of estuarine open water area, based on data in Eyre et al., 2011 and Zhao et al., 2008 f. From Pendleton et al., 2012 g. Seagrass N2O fluxes estimated based on N2O/N ratios of 0.001 to 0.06, from Nishio et al., 1983, Jensen et al., 1984, and Seitzinger, 1988
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Max
Accepted Article
Environment
Global Flux Estimate (Tg N2O-N yr-1)
Reference
Mangrove
0.01 - 0.09
This study
Mangrove
0.014 - 0.56
Corredor et al., 1999
Mangrove
0.11
Chauhan et al., 2008
Mangrove
0.07
Barnes et al., 2006
Salt marsh
0.001 - 0.14
This study
Intertidal sediment
0.007 - 0.05
This study
Seagrass
0 - 0.19
This study
Estuarine open water
0.13 - 0.44
This study
Estuarine open water
0.13 - 0.45
Robinson et al., 1998
Estuarine open water
0.22
Law et al., 1992
Estuarine open water
1.5
De Wilde and De Bie (2000)
Estuarine open water
3.7 - 5.7
Bange et al., 1996
Estuarine open water
0.07 - 0.69
Seitzinger and Kroeze, 1998
Estuarine open water Total of estuarine environments Rivers and estuaries combined
0.25
Kroeze et al., 2005
0.17 - 0.95
This study
0.6
Syakila et al., 2010
Table 7. Estimated global N2O fluxes from previous studies.
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Accepted Article This article is protected by copyright. All rights reserved.
Accepted Article This article is protected by copyright. All rights reserved.
Accepted Article This article is protected by copyright. All rights reserved.
Accepted Article This article is protected by copyright. All rights reserved.