Atmospheric nitrate export in streams along a montane to urban gradient
Graphical abstract
Introduction
Atmospheric nitrogen (N) deposition has increased 10-fold over the past century, increasingly contributing to the global N availability (Galloway et al., 2004). Anthropogenic activities such as fossil-fuel combustion, agriculture, and fertilizers use are responsible for this increase (Fowler et al., 2015; Galloway et al., 2008; Vitousek et al., 1997), with impacts observed in remote ecosystems (Hastings et al., 2009; Holtgrieve et al., 2011; Preunkert, 2003). High N loading to the environment has been documented for many ecosystems (Aber et al., 1989; Clark et al., 2017; Elser et al., 2009; Matson et al., 2002). To address this issue, global efforts are underway to alleviate N inputs into ecosystems, aiming at “minimizing the consequent harm to humans and the environment” (International Nitrogen Initiative, http://www.initrogen.org). Nitrate (NO3−) concentrations, and fluxes in soils and streams, have often been used to assess the N saturation status in watersheds (Aber et al., 1989; Baron and Campbell, 1997; Lovett and Goodale, 2011). However, N exports in streams depend on multiple parameters such as basin topography (Balestrini et al., 2013; Clow and Sueker, 2000), land-cover (Barnes et al., 2014; Williams et al., 2016) and land-management (Barnes and Raymond, 2010; Burns et al., 2009; Lefebvre et al., 2007). As streams integrate many processes at the watershed scale, understanding the specific sources of NO3− is necessary to (1) evaluate the respective contribution of natural and anthropogenic sources, (2) clarify the fate of deposited N in the environment, and (3) understand the origin and the development of N saturation.
Mountainous ecosystems are particularly sensitive to increased N inputs by atmospheric deposition (Baron et al., 2000, Baron et al., 2005, Baron et al., 2011), as they are historically N-limited (Kaye and Hart, 1997). Critical N loads for these ecosystems are among the lowest for pristine environments (Baron et al., 2011; Bowman et al., 2006; Nanus et al., 2017), making them vulnerable to long-range transport of atmospheric N emitted from distant sources (Mast et al., 2014; Wasiuta et al., 2015). Atmospheric deposition of N has been shown to contribute significantly, either directly or indirectly, to year-round NO3− exports from mountainous catchments (Hundey et al., 2016; Nanus et al., 2008), typically showing a pulse at spring as soils subsurface NO3− reservoirs are flushed by snowmelt water (Kendall et al., 1995; Williams et al., 2009; Williams and Melack, 1991).
Atmospheric deposition is also a major source of N to urban areas, which receive much higher loads than adjacent environments (Bettez and Groffman, 2013; Fang et al., 2011; Hall et al., 2014; Rao et al., 2014; Templer et al., 2015). Local and regional emissions and subsequent deposition of fuel-combustion derived NOx and ammonia (NH3) are responsible for this pattern (Galloway et al., 2004; Kean et al., 2000). NOx compounds are oxidized into NO3− within hours (Beirle et al., 2011), then scavenged from the atmosphere by wet and dry deposition (Hertel et al., 2012). In the atmosphere, NH3 is in equilibrium with ammonium (NH4+), the other primary component of bulk N deposition. Because of the particular topography of urban basins (extended impervious surface, rapid precipitation runoff), urbanization can lead to high NO3− exports in freshwater bodies (Groffman et al., 2004; Riha et al., 2014), with major ecological, economic and health consequences (Dodds et al., 2009).
A number of previous studies have used the dual isotope approach (δ18O and δ15N of NO3−) to track the spatio-temporal variability of sources contribution to NO3− pools in a large variety of environmental matrixes (Campbell et al., 2002; Durka et al., 1994; Elliott et al., 2009; Yang and Toor, 2016). Biochemical processes such as denitrification or assimilation have also been shown in laboratory experiments to distinguishably enrich residual NO3− in heavier isotopes, here 15N and 18O (Granger et al., 2004, Granger et al., 2010; Treibergs and Granger, 2017), although this enrichment can be diluted by newly nitrified NO3− with low N and O isotopic values (Granger and Wankel, 2016; Mayer et al., 2002). Environmental studies also reported characteristic NO3− isotopic enrichment for denitrification (Clément et al., 2003; Fang et al., 2015; Wexler et al., 2014), assimilation (Emmerton et al., 2001; Estrada et al., 2017; Liu et al., 2013b) or photolysis (Frey et al., 2009; Shi et al., 2015; Ye et al., 2016). However, the isotopic fingerprint of biological processes can lead to inaccurate NO3− source apportionment in some cases, especially in delineating the respective contribution of the microbial and the atmospheric sources (Michalski et al., 2004; Riha et al., 2014; Rose et al., 2015b). In the past few years, a growing number of studies have used an isotopic particularity of NO3−atm to quantify the contribution of atmospheric deposition to terrestrial N pools (Costa et al., 2011; Hundey et al., 2016; Tsunogai et al., 2014). NO3−atm is enriched in 17O due to its production pathways (i.e., oxidation of NOx by O3), showing a deviation from the Terrestrial Fractionation Line (Thiemens, 2006). Δ17O is a quantification of this deviation, calculated as Δ17O = δ17O − 0.52 ∗ δ18O in the present work. Δ17O value of NO3−atm generally ranges between 20 and 35‰ in temperate latitudes (Morin et al., 2009; Savarino et al., 2007), whereas Δ17O value of NO3− from all other sources (industrial fertilizers, nitrification) or of biologically processed NO3−atm, is 0‰ (Michalski et al., 2004, Michalski et al., 2015). Because NO3− loss processes (i.e., denitrification, assimilation or photolysis) obey the mass-dependent fractionation law, Δ17O-NO3− can be used as a conservative tracer of NO3−atm in the environment and help to better estimate NO3−atm contribution to streams NO3− pool (Michalski et al., 2004).
This study evaluates how atmospheric deposition of N contributes to NO3− exports in several streams along a montane to urban gradient in the French Alps. Isotopic (Δ17O, δ15N, δ18O of nitrate, δ2H and δ18O of water) and in situ hydro-chemical techniques were combined to evaluate the drivers of NO3− inputs and removal in all streams. The first hypothesis was that due to a higher contribution from other sources (e.g., sewage, fertilizers) in urban areas compared to the mountains, coupled with higher local atmospheric N inputs, total NO3− and atmospheric nitrate (NO3−atm) exports should increase along the gradient. The second hypothesis was that different NO3−atm export dynamics should be observed in urban streams compare to montane streams, due to different hydrological drivers that need to be determined. To test these hypotheses, (1) NO3− concentration and its isotopic composition were determined in six streams and one reservoir, ranging from 2000 m above sea level (a.s.l.) to 200 m (a.s.l.) in the French Alps, and the annual export fluxes of NO3−atm and total NO3− were compared across sites along the gradient, (2) the drivers of the seasonal variability in NO3−atm proportion were identified and (3) the additional sources of NO3− in each watershed were determined.
Section snippets
Study site and selected streams
The Romanche Valley, located in the central French Alps, spreads from the Lautaret pass (2058 m a.s.l.) down to Grenoble (250 m a.s.l.) (Fig. S1). The Grenoble conurbation counts around 500,000 inhabitants, and is the biggest alpine metropolis in France.
Six streams and one reservoir were sampled from the Lautaret pass to Grenoble, draining watersheds with distinct geomorphic and biogeographic characteristics (Table 1). Two alpine streams were sampled at ~2000 m a.s.l.. These drain the South exposed
Hydrological patterns
Cumulated precipitation was 537 and 609 mm in 2015 and 804 and 740 mm in 2016 at the upper montane and the urban sites, respectively. Discharge at the lower montane site reflects a snowmelt influenced hydrological regime. It peaked up to 30 m3 s−1 in spring, and was significantly higher from April to October compared to the rest of the year. Discharge at both urban sites was consistent with a hydrological regime driven by snowmelt in spring-summer (main maximum), and rainfall in autumn (secondary
Total NO3− and NO3−atm exports in streams
There were little differences in stream [NO3−] along the elevation gradient despite an expected higher contribution from anthropogenic activities to streams N pool closer to urbanized areas (Groffman et al., 2004; Tsunogai et al., 2016). The 0.2–1.4 mg L−1 mean [NO3−] range at the montane sites is consistent with the 0.6–1.6 mg L−1 range in other Alpine valleys (reviewed in Balestrini et al., 2013), and the 0.9–1.5 mg L−1 range in elevated sites of the Colorado Front Range (Baron and Campbell, 1997;
Conclusions
The results showing higher NO3−atm exports in a montane stream relative to urban streams have several key implications. First, it sheds light on the contribution of snowmelt-derived groundwater to year-round NO3−atm exports in all streams at baseflow. Contamination of groundwater by NO3−atm may be a widespread phenomenon, regardless of emitting sources proximity: other studies also reported ubiquitous presence of NO3−atm in groundwater (Dejwakh et al., 2012; Dietzel et al., 2014; Nakagawa et
Acknowledgments
This study was supported by grants from the Labex OSUG@2020 (“Investissements d'avenir” - ANR10 LABX56), the ARC – Environnement Région Rhone-Alpes, the Grenoble-Chambéry DIPEE CNRS. This work also benefited from the National Research Agency supports (“Investissements d'avenir” - ANR11 INBS-0001AnaEE -Services and “FloodScale project” - ANR 2011 BS56 027) and from the SAJF research station (UMS 3370, UGA-CNRS) infrastructures and competences. The study took place on a Long Term Ecological
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