Current atmospheric nitrogen deposition still exceeds critical loads for sensitive, semi-natural ecosystems in Switzerland

Atmospheric Environment 211 (2019) 214–225

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Current atmospheric nitrogen deposition still exceeds critical loads for sensitive, semi-natural ecosystems in Switzerland

T

Zaida Kosonena,∗, Elvira Schnydera, Erika Hiltbrunnerb, Anne Thimonierc, Maria Schmittc, Eva Seitlera, Lotti Thönia FUB – Research Group for Environmental Monitoring, Alte Jonastrasse 83, 8640, Rapperswil, Switzerland Department of Environmental Sciences, Institute of Botany, University of Basel, Schönbeinstrasse 6, 4056, Basel, Switzerland c Swiss Federal Institute for Forest, Snow and Landscape Research WSL, Zürcherstrasse 111, 8903, Birmensdorf, Switzerland a

b

A R T I C LE I N FO

A B S T R A C T

Keywords: N load Inferential method Grassland Forest Wetland Temporal development

Increased atmospheric nitrogen (N) deposition is driving nutrient imbalances, soil acidification, biodiversity losses and the long-term reduction in stability of sensitive ecosystems which previously had limited N. In this study, we analysed the concentrations of seven different N compounds in precipitation and in the air at 34 sites across Switzerland. We calculated the N deposition by precipitation (bulk deposition) and applied the inferential method to derive dry deposition (gases, aerosols) from air concentrations. We then quantified the total inorganic N deposition by adding together the bulk and dry deposition. Finally, the total inorganic N input into the sensitive ecosystems of the 34 sites was compared to the critical loads of these ecosystems. N deposition by precipitation was the main contributor to the total N load in 16 out of 34 sites, especially into open ecosystems such as alpine/subalpine grassland, mountain hay meadows, and raised bogs. Dry deposition of ammonia (NH3) was the second most important pathway, in particular for forests close to agricultural activities, due to high NH3 concentrations and the higher deposition velocity. The N deposition exceeded the lower limit of the Critical Load of Nitrogen (CLN) range at most sites, and at many sites even surpassed the upper limit of the CLN range. No, or minor, exceedances of the critical loads for N were found only at remote sites at higher elevation in the Central Alps. Annual inorganic N deposition between 2000 and 2017 revealed a significant decline in oxidised N compounds at four of five sites (−1.6–1.8% per year), but reduced compounds only decreased at two sites (−1% and −1.4% per year) and even increased at one site (+1.2% per year), despite adopted abatement strategies for agricultural practices. This emphasises that most sensitive ecosystems in Switzerland continue to be exposed to excessive N loads through atmospheric deposition, with detrimental consequences for the biodiversity and stability of these ecosystems.

1. Introduction Until several decades ago, in most ecosystems in Europe and elsewhere, the growth of plants was often limited by nitrogen (N) availability (Fay et al., 2015; Vitousek et al., 2010). N emissions have steadily risen and N deposition into sensitive ecosystems (forests, bogs, and mountain hay meadows) has been increasing since the 1950s, causing over-fertilisation, nutrient imbalances, and soil acidification. This has negative consequences on biodiversity and the stability of ecosystems (Aber et al., 1995; Berendse et al., 1993; Bobbink et al., 2010; Fowler et al., 2013; Galloway et al., 2008; Krupa, 2003; Tilman, 1987). Regulatory measures for reducing N emission have been implemented in order to protect human health and sensitive ecosystems:

∗

In Switzerland, catalytic converters in cars have been in use since the 1980s (NOx reduction; OAPC, 1985), flue gas denitrification in industrial settings has been established, and measures have been recommended to the agricultural sector such as the use of low emission application methods for slurry, and the use of closed tanks for liquid manure storage (Biedermann et al., 1996; BUWAL, 2003; FOEN, 2015). However, several of these measures are not mandatory, or have to be applied only when new agricultural buildings have been set-up. To determine the effectiveness of all the measures taken for the reduction of N emissions and thus atmospheric deposition, efficiency evaluations should recurrently be carried out by quantifying the total deposition of N. The total deposition is compared to the critical loads for N for the corresponding ecosystem. Critical loads for N (CLN) are

Corresponding author. E-mail address: [email protected] (Z. Kosonen).

https://doi.org/10.1016/j.atmosenv.2019.05.005 Received 11 December 2018; Received in revised form 26 March 2019; Accepted 4 May 2019 Available online 12 May 2019 1352-2310/ © 2019 Elsevier Ltd. All rights reserved.

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defined as “a quantitative estimate of an exposure to [nitrogen] below which significant harmful effects on specified sensitive elements of the environment do not occur according to present knowledge” (UNECE, 2012 based on Nilsson and Grennfelt, 1988). Since critical load exceedances are not necessarily associated with immediate negative effects, but rather the avoidance of negative long-term effects, efforts to reduce emissions must continue. The Critical Loads for Nitrogen used in this study have been determined empirically for various ecosystem types, largely based on N addition experiments in the field and in mesocosms (reviewed and summarized in Bobbink and Hettelingh, 2011). Different inorganic N compound in the atmosphere contribute to the total atmospheric N input into sensitive ecosystems: dissolved ammonium (NH4+) and nitrate ions (NO3−) in precipitation; the gases ammonia (NH3), nitrogen dioxide (NO2), and nitric acid (HNO3); and NH4+ and NO3− in aerosols. These seven compounds have been quantified in the present study. The contribution of each compound to the total inorganic N deposition varies with different geographical and meteorological properties such as precipitation, elevation, and proximity to sources. Switzerland is divided into five biogeographic regions which differ in these properties and therefore we investigated the contribution of each N compound within each region. The different N compounds are either primary components directly resulting from a N source, such as NH3, or they are secondary components resulting from chemical reactions in the atmosphere. The N compounds: nitrous acid (HNO2), nitric oxide (NO) and peroxyacetyl nitrate (PAN) were not included here, as they are unstable and/or they react into components that are accounted for by our methods (Singh, 1987; Singh et al., 1986). The two major pathways of atmospheric deposition are wet and dry deposition. A third pathway, occult deposition, is deposition in the form of fog (often neglected because of collecting constraints), and will not be addressed here. Wet deposition is defined as the deposition of NH4+ and NO3− in the form of rain, hail, and snow. It is measured using collecting devices that open only during precipitation events (wet-only collector). For practical reasons, however, precipitation is often sampled using continuously open collectors (bulk collectors), as is the case in our study. The input quantified by such devices is called “bulk” deposition, which includes the gravitational particulates deposited on the collecting surfaces during dry periods. In the bulk deposition, an unknown amount of dry deposition originating from gases and aerosols is captured along with the wet deposition. However, it can be expected that the amount of dry deposition collected on the smooth surface of the collecting device is smaller than on the rough surfaces of forests and grasslands. Furthermore, the narrow opening of the funnel is expected to only allow for a small fraction of gases and aerosols to be dissolved. Dry deposition is defined as the deposition of gases and aerosols. It is considerably more complex to determine dry deposition than bulk deposition. The most common method used to assess dry deposition is the inferential method (Schmitt et al., 2005). For quantifying the dry deposition, concentrations of gasses and aerosols in the air at the study site (in an open area) are determined and multiplied by deposition velocities (vd). Deposition velocities depend on the N compound and on the surface where it is deposited. A wide range of deposition velocities are found in the literature (Schrader and Brümmer, 2014). For the present study, we used the yearly averaged vd recommended by Rihm and Achermann (2016) and took into account the duration of snow cover where necessary. The vd for NH3, NO2 and aerosol accounted for different vegetation types with highest values for conifer forests and lower values for low-stature vegetation. These vd values have been defined and commonly used for monitoring in Switzerland, thus ensuring that our analysis will be comparable to previous work. In our study, we define the total deposition of inorganic N as the sum of bulk deposition (wet deposition with inclusion of gravitational dust) and dry deposition of gaseous NH3, NO2, HNO3 and NH4+ and NO3− in aerosols. The methods used in this study are easy to use and inexpensive for

assessing different N compounds. These methods do not account for all nitrogen compounds nor all mechanisms of N deposition such as, for example, the bidirectionality (storing and release) of nitrogen via dew (Wentworth et al., 2016). However, these methods are designed for simple monitoring purposes, and what is lost in precision is gained by the ability to asses wider and/or denser monitoring networks. Regardless of the tradeoffs, these methods have been shown to give accurate estimations of N input into ecosystems and are exact enough for monitoring the deposition of inorganic N in compliance with the CLN. The objectives of this study are to (i) demonstrate the extent of the variation in the total inorganic N deposition and the contribution of the different N components to the total N deposition across the five biogeographic regions in Switzerland, (ii) quantify the current atmospheric deposition of N into various sensitive ecosystems in relation to CLN loads, and (iii) present the temporal trends of N deposition between 2000 and 2017. 2. Material and methods 2.1. Study sites The main measuring campaign of the atmospheric inorganic N deposition in Switzerland was carried out in 2014, on 34 long-term monitoring sites of various institutions (Table S1). All sites were located in areas of different land use types and within, or close to (< 200m), different sensitive ecosystem types such as bogs, forests, and (sub-)alpine grasslands (Table 1). The information about ecosystem types is based on the Geo-catalogue of the Federal Office of Topography (map.geo.admin.ch). The study sites were distributed across the five different biogeographic regions of Switzerland (Jura, Plateau, Northern, Central, and Southern Alps; Fig. 1, adapted from Gutersohn, 1973). These regions are based on the distribution patterns of flora and fauna and differ in climatic properties, human population density, traffic volume as well as in the intensity of agricultural activities. Fourteen sites were run by WSL (Swiss Federal Institute for Forest, Snow and Landscape Research) within the framework of the long-term forest ecosystem research programme LWF (www.wsl.ch/lwf), which is part of the UNECE International Co-operative Programme on Assessment and Monitoring of Air Pollution Effects on Forests (ICP Forests; www.icp-forests.net) and the LTER-Europe Network (LongTerm Ecosystem Research in Europe; https://lternet.edu/). ANA B is a separate project of the WSL, aiming to assess the impact of experimentally elevated N deposition (Schleppi et al., 2017). These 15 sites are referred to as WSL sites. Three sites are part of the NABEL stations, the Swiss National Air Pollution Monitoring Network (www.empa.ch/ web/s503/nabel); named NABEL sites. TIE, FUR and KLE are sites of the Alpine Research Station Furka (ALPFOR; www.alpfor.ch), ARD 02 und 06 are sites run by Agroscope (www.agroscope.ch), and MU 01 and ZB 01 are run by the Institute for Applied Plant Biology (IAP) within an Intercantonal Forest Observation Programme (Braun et al., 2018). The other stations belong to the ammonia monitoring network in Switzerland of several cantons and the FOEN (Seitler et al., 2018). The measurements (bulk deposition, NH3, NO2) at the ALPFOR, Agroscope, and cantonal sites were conducted by the FUB (Research Group for Environmental Monitoring); named FUB sites hereafter (Seitler et al., 2016; Table S1). FUB further carried out the NH3 measurements at the three NABEL sites, and the aerosols measurements at CHA. Changes in the N deposition between 2000 or 2002 and 2017 were assessed at five sites including one bog and four forest sites (MAG, SCH, BA, RIG, CHA). 2.2. Field measurements Depending on the site, the collectors/samplers were exposed for periods of either one week, two weeks, four weeks, or one month during the year 2014 or – at the sites with long-term measurements of all N 215

Bettlachstock Chaumont Tourbières, NE Nenzlingen Bachtel Hudelmoos Jussy Lägeren Lausanne

Lengwiler Weiher Muri 1 Othmarsingen Rothenturmer Hochmoor Vordemwald

Wauwiler Moos Alpthal Beatenberg Klewenalp Rigi-Seebodenalp Schänis Sörenberg Husegg Zugerberg 1 Ardez Szen. 2 Ardez Szen. 6 Celerina Davos Furka Nationalpark Tiefenbach Visp Chironico Magadino-Cadenazzo Novaggio Sagno Reservoir

BET CHA NE 01 NENZ BA HUD JUS LAE F LAU

LEN MU 01 OTH ROHO

WAU ANA B BEA KLE RIG SCH SOER ZB 01 ARD 02 ARD 06 CEL DAV FUR NAT TIE VIS CHI MAG NOV SARE

216

P NA NA NA NA NA NA NA CA CA CA CA CA CA CA CA SA SA SA SA

P

P P P P

J J J J P P P P P

Region

8.01 8.71 7.77 8.48 8.46 9.06 8.01 8.53 10.18 10.17 9.88 9.86 8.42 10.24 8.46 7.86 8.82 8.93 8.84 9.04

7.91

9.18 8.36 8.22 8.68

47.17 47.04 46.7 46.93 47.07 47.15 46.81 47.13 46.79 46.78 46.51 46.81 46.58 46.66 46.59 46.3 46.44 46.16 46.02 45.86

47.27

47.63 47.27 47.4 47.12

47.22 47.05 46.99 47.45 47.3 47.52 46.23 47.47 46.56

°

°

7.41 6.98 6.74 7.57 8.9 9.29 6.29 8.35 6.65

N

E

Coordinates

499 1190 1560 1722 1031 630 1450 990 2180 1680 1760 1629 2440 1900 2127 664 1479 204 1075 820

486

505 470 462 905

1076 1137 1000 500 930 520 501 508 790

m a.s.l.

Elevation

FUB WSL WSL FUB NABEL WSL FUB FUB FUB FUB WSL WSL FUB WSL FUB WSL WSL NABEL WSL FUB

WSL

FUB FUB WSL FUB

WSL1 NABEL2 FUB3 FUB FUB FUB WSL WSL WSL

N in precipitation measured by

1 WSL = Swiss Federal Institute for Forest, Snow and Landscape Research WSL. 2 NABEL= The Swiss Federal Laboratories for Materials Science and Technology (Empa). 3 FUB = Research Group for Environmental Monitoring. a snow cover was taken into account for the evaluation of the vd. b Precipitation measured at nearby weather station.

VOR

Site Name

Site Code

Meadow, intensive agriculture Intensive agriculture No agricultural use Pasture Pasture Meadow, pasture Pasture Pasture Meadow, pasture Pasture Meadow, pasture No agricultural use No agricultural use Pasture No agricultural use Pasture No agricultural use Pasture Intensive agriculture Pasture No agricultural use

Meadow, pasture Meadow, pasture Meadow, pasture Meadow Meadow, pasture Meadow, pasture Meadow, pasture Meadow, pasture Meadow, intensive agriculture No agricultural use No agricultural use No agricultural use No agricultural use

Land use surrounding site

bog bog alpine/subalpine alpine/subalpine mixed forest coniferous forest raised bog raised bog alpine/subalpine alpine/subalpine raised bog raised bog alpine/subalpine alpine/subalpine alpine/subalpine coniferous forest alpine/subalpine bog alpine/subalpine mixed forest

mixed forest

bog mixed forest mixed forest raised bog

grassland

grassland

grassland grassland grassland

grassland grassland

grassland grassland

alpine/subalpine grassland mixed forest raised bog mountain hay meadow mixed forest raised bog deciduous forest mixed forest deciduous forest

Sensitive ecosystem type (within 200 m)

884 2183 1228 1228 1523 1349 1893 1349 756 691 920 993 1280b 967 1087 656 1827 2592 2558 1903

1018

746 930 896 1635

1209 1054 1183 890 1718 904 942 1015 1206

mm a

−1

Precipitation 2014

−1

10–15 10–15 5–10 5–10 5–20 5–15 5–10 5–10 5–10 5–10 5–10 5–10 5–10 5–10 5–10 5–10 5–10 10–15 5–10 5–20

5–20

10–15 5–20 5–20 5–10

5–10 5–20 5–10 10–15 5–20 5–10 10–20 5–20 10–20

kg h a−1

CLN

2 2 1.5 1.5 3.5 4 2 2 1.5 1.5 2 2 1.5 1.5 1.5 4 1.5 2 1.5 3.5

3.5

2 3.5 3.5 2

1.5 3.5 2 1.5 3.5 2 3 3.5 3

mm s

NO2

vd

−1

20 19a 11* 10a 26 30 15a 18a 9a 10a 18* 16* 8a 10a 10* 30 11a 20 12 26

26

20 26 26 17*

11* 26 18* 12 26 20 22 26 22

mm s

NH3 −1

1.25 2 2 2 4 4 2 2 2 2 2 2 2 2 2 4 2 2 2 4

2.4

1.25 2.4 2.3 2

2 4 2 1.25 4 1.25 1.9 2.5 3

mm s−1

Aerosol

Table 1 Study sites (sorted by region and alphabetically) including surrounding land use and sensitive ecosystem type with CLN and vd used for each site. CLN is based on Bobbink and Hettelingh (2011) and CLN mountain hay meadows revised after Roth et al. (2013). The vd are based on Rihm and Achermann (2016) and Schmitt et al. (2005). Regions of Switzerland: J = Jura, P=Plateau, NA=Northern Alps, CA=Central Alps, SA=Southern Alps. Pastures are grazed, meadows are cut for hay making.

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Fig. 1. The selected 34 sites in different regions of Switzerland (J = Jura, P = Plateau, NA = Northern Alps, CA = Central Alps, SA = Southern Alps). Symbols represent different sensitive ecosystem types, a/s grassland refers to alpine/sub-alpine grassland.

compounds considered in this study (BA, SCH, CHA, MAG, and RIG) – from 2000/2002 on.

Supplementary Tables S1 and S2).

2.2.2.1. Passive sampler for NH3. From 2003 onwards, the concentration of NH3 in the air was measured with Radiello samplers (Fondazione Salvatore Maugeri, Italy) in accordance to VDI 3869 Part 4, 2012. The Radiello passive sampler is a cylindrical, coaxial system with a cartridge as an absorption medium and a diffusive uptake rate of NH3 that is at least three times higher than any axial diffusive sampler (http://www.radiello.com/english/nh3_en.htm). For quality control purposes, two passive samplers were used at each site, placed under a rain shelter. For the temporal data, the “Zürcher” passive sampler was used until 2003 (described in Thöni et al., 2003). To ensure the comparability of data from both types of samplers, a multi-year comparison of the two samplers’ measurements was carried out (November 2002–June 2006, ten sites over one to four years). The values obtained by the Zürcher samplers were slightly lower than those obtained by the Radiello samplers and so were corrected by a factor of 1.1.

2.2.1. Precipitation (bulk and wet deposition) The method to measure inorganic N deposition used at the WSL sites (except ANA B) is described in detail in Thimonier et al. (2005). Bulk precipitation was collected by three 2-L polyethylene (PE) bottles connected to funnels with a 100-cm2 opening. At the WSL sites where snowfall was expected (BEA, BET, CEL, CHI, DAV, NAT, NOV), the funnel-type collector was replaced by a single bucket-type collector (30 cm diameter) during wintertime. Similar collectors were used for the ANA B site, with two funnel-type collectors (500 cm2 opening) during the summer, and a snow bucket (1000 cm2 opening) during the winter (Schleppi et al., 2017, 1998). The FUB bulk collector consisted of two 5-L PE bottles, one upsidedown and without a bottom functioning as a funnel, (196 cm2 opening) bolted together over the bottleneck (method following VDI 4320 Part 3, 2017; further details in Kosonen et al., 2018). To ensure the comparability of the bulk collectors (funnels and snow buckets), the WSL and the FUB laboratories performed a comparison of the sampling and analysis methods that were applied in the field. The intercomparison yielded very similar results for the NH4+ and NO3− concentrations in the bulk samples (Seitler et al., 2016 and Kosonen et al., 2018, Table S3). Wet-only collector (Type DRA-12HK, Digitel, Switzerland) were used at the three NABEL sites (Empa, 2015). We corrected the wet-only deposition data as our measurements were based on bulk collectors. Comparisons of the corrected wet-only and bulk deposition revealed a significant linear correlation (see Table S4). Wet-only values were multiplied by a factor of 1.15 for NH4 and by 1.17 for NO3−.

2.2.2.2. NO2 measurements. For measuring NO2, a Palmes-type sampler (Palmes and Gunnison, 1973) was used at most sites. It consisted of a PMMA (polymethylmethacrylate) tube with a polyethylene cap and three stainless steel grids dipped and coated in 25% (v/v) triethanolamine/acetone solution. A PTFE (polytetrafluorethylene) membrane was placed at the inlet as a wind shelter to keep the diffusion path constant (Brunner and Schlatter, 2002). Two passive samplers were placed under the same shelter as the Radiello samplers. Specific temperature and pressure are known to affect the gas diffusivity, but this meteorological data was not available for our sites. However, no large deviation in the annual mean was noticed between reference methods and the NH3 and NO2 concentrations gained by the passive samplers (see Table S5). In addition, temperatures of plant canopies largely differ from air temperatures, especially in low-stature grassland (Körner and Hiltbrunner, 2018), thus, adjustment by air temperature does not simulate more realistic deposition conditions. At the three NABEL sites, NO2 was measured by a continuous measurement system based on chemoluminescence (Thermo 42i-TL, Thermo Fisher Scientific,USA).

2.2.2. Air concentrations of nitrogen compounds (inferential method) Air concentrations of NH3 and NO2 were measured using passive samplers except at the three NABEL sites, where NO2 concentrations were quantified continuously by chemiluminescence (DIN EN 14211:2012–11). HNO3, aerosol-NH4+, and aerosol-NO3- concentrations in the air were measured at 17 sites (including 8 sites which were not a part of this study) using denuders and filters. Based on these values, the concentrations were estimated for all other sites (see 217

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2.2.2.3. Denuder filter systems (active). Air was pumped through denuders and subsequently through filters connected to the denuders (VDI 3869 Part 3, 2010; self-made equipment). For measuring gaseous HNO3, the coating was KOH (2% g/v), for aerosol-NO3- NaCl (1% g/v) and for aerosol-NH4+ oxalic acid (2% g/v), respectively. The air pumps were type PM 16034-NMP05 (KNF Neuberger AG, Switzerland) with flow rates of 6–8 m3 within 14 days. In 2014 denuders were installed at eight sites only. For the rest of the sites estimated values based on the most similar sites were used instead (see Table S1). According to HNO3 concentration maps generated by methods published by ICP Materials (UNECE, 2005) and to maps of secondary aerosols based on the calculation offered by Roth et al. (2013), these three compounds do not deviate much even for a larger region.

(https://www.nilu.no/projects/ccc/intercomparison/index.html, Table S6). All analytical results compared within the QA were found to be in good agreement (see Table S7, Table S8, Table S9). The results from the deposition sample analysis of the WSL sites were also regularly checked by comparing measured and calculated electrical conductivity, calculating the ion balance (König et al., 2013). Analyses were repeated if suspicious values were detected or variability of duplicate samples was higher than 10%. The quality assurance and control procedures further included the use of control charts for internal reference material and since 2002 the WSL laboratory regularly participates in periodic working ring tests (e.g., Marchetto et al., 2011). The NABEL samples were regularly checked by comparing measured and calculated electrical conductivity, and by calculating the ion balance. The reliability of the laboratory analysis has been ensured by yearly ring tests (https://www.nilu.no/projects/ccc/intercomparison/ index.html).

2.3. Chemical analyses 2.3.1. Precipitation The samples collected at the WSL sites were filtered (0.45 μm) and stored for a few weeks at most at 2 °C before being analysed. NO3− was measured by ion chromatography (IC; DX-120 Dionex, Thermo Fisher Scientific, USA), and NH4+ was measured by flow injection analysis (FIAS 300, PerkinElmer, USA). Samples of the FUB collector were stored at 4 °C and analysed within a few days, or stored frozen (−18 °C) before further analysis by the FUB laboratory. NH4+ was analysed by a flow injection analyser (FIA; Foss FIAstar 5000, Sweden, with a gas diffusion membrane, detection by UV/VIS photometry; SN EN ISO 11732:2005), and NO3− by ion chromatography (IC; ICS-1600 Dionex, Thermo Fischer Scientific, USA, DIN EN ISO 10304–1: 2009–07). The samples of the wet-only deposition (NABEL) were stored at 4 °C or stored frozen (−20 °C). NH4+ was analysed with ion chromatography (IC) Dionex CG12A and CS12A, and NO3- with IC Dionex AG14A and AS14A (Thermo Fischer Scientific, USA). Further details can be found in Empa (2015).

2.4. Estimation of total inorganic nitrogen deposition 2.4.1. Precipitation Fluxes of dissolved NH4+ and NO3− were obtained by multiplying the measured concentrations with the amount of precipitation per sampling period and by cumulating these fluxes over the whole sampling year. The yearly concentrations were calculated as weighted averages. The precipitation amounts of the bulk collectors from the sites at the highest elevation in the Central Alps (FUR, TIE) and in the Northern Alps (KLE) were compared with the measured precipitation amount at the weather station nearby, installed ca. three meters from the bulk collectors (FUR: www.alpfor.ch and TIE: www.monitron.ch). At the KLE site, an additional rain tipping gauge was installed as the next weather station was located at a distance of 430 m to KLE. In particular, at the FUR site, precipitation amounts of the bulk collectors were consistently lower than the amounts measured by the nearby weather station (on average −21% for the whole year 2014). In cases when the precipitation fell as snow, the snow was largely blown out of the bulk collector at this site. Therefore, precipitation amounts were corrected for FUR. For two further high-elevation sites (KLE, TIE) where there was a weather station, the amounts in the bulk collectors were not consistently lower than the weather station or the tipping gauge (KLE), respectively, thus bulk precipitation amounts were not adjusted.

2.3.2. Passive samplers (PS) and denuder NH3, NO2, HNO3, NH4 and NO3 were all measured by FUB (exception: all compounds of RIG and MAG and NO2–N of CHA by NABEL). The Radiello cartridges were eluted with ultrapure water and NH3 was measured like NH4+ with FIA (see above). The concentration in air was determined in accordance with the instructions provided by the Radiello manufacturer. The solution of the Zürcher PS was analysed by ion chromatography (Dionex DX-100, Dionex column CS 12, Thermo Fisher Scientific, USA). The NO2− absorbed on the coated grids was eluted with an aqueous solution consisting of sulphanilamide (0.06 mol L−1), phosphoric acid (0.37 mol L−1), and N-(1-)Naphtyl(ethylendiamin-dihydrochloridmonomethanolat (NEDA, 0.24 mmol L−1), quantified by the GriessSaltzman reaction and analysed by UV/Vis spectrophotometry in the visible range (at 540 nm). The diffusion coefficient used for determining the air concentration was 1.53 × 10−5 m2 s−1. HNO3 and aerosol-NO3- of FUB samples were measured with IC, aerosol-NH4+ were measured with FIA (FUB), and samples from the NABEL sites (RIG und MAG) were measured with IC (see above analysing procedure of precipitation).

2.4.2. Inferential method The total deposition of N was obtained by summing the dry deposition of all five gas or aerosol compounds, and then adding the bulk deposition of NH4+ and NO3−. The annual dry N deposition from gaseous NH3, NO2, HNO3, and from NH4+ respectively NO3− in aerosols was obtained by multiplying the annual air concentrations of each N compound by the corresponding average annual deposition velocity (vd; see Schmitt et al., 2005 for a detailed description). The vd in this study were taken from Rihm and Achermann (2016). For NO2, NH3 and aerosols, vd are dependent on vegetation roughness, the duration of snow cover in nonforested ecosystems (NH3) and for aerosols on the altitude. Thus, the vd applied here differed between the selected sites and between the ecosystem types (Table 1). We used a constant value of 15 mm s−1 for vd of HNO3 as for Switzerland a constant vd has been defined.

2.3.3. Quality assurance (QA) Blank and duplicate samples (bulk, NH3, NO2) were used by the FUB laboratory throughout all processes in order to assess material and analysis quality as well as to detect contamination during sampling and analysis procedures. Values of HNO3-denuders and aerosol-NO3- were indirectly verified by comparing them to filters that quantified the sum of both compounds. Furthermore, the results were compared with other independent methods and/or other laboratories and certified reference material was used whenever available. The reliability of the laboratory analysis has been ensured by yearly inter-laboratory comparisons

2.5. Critical loads for nitrogen CLN is defined separately for each sensitive ecosystem type and is indicated as a range with a lower and upper limit (Bobbink and Hettelingh, 2011; Roth et al., 2013). For the 34 sites, the ecosystem type and the CLN ranges are listed in Table 1. The measured inorganic N deposition values were compared with the CLN of the sensitive ecotypes. The main sensitive ecosystem type on-site or close to the 218

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were similar but due to lower precipitation in the Jura, the deposition was also lower (median annual deposition in kg ha−1 NA: 4.2, J: 2.6).

measurement site (within 200 m) was included. 2.6. Statistical analysis

3.1.2. Dry inorganic nitrogen deposition The deposition of NH3–N contributed considerably to the total inorganic N deposition (7–65%, median 33%), whereas the input of NO2–N was only minor (2–24%, median 7%) (Fig. 2). Overall, the NH3–N concentration was lower than the NO2–N concentration at most sites (Table 2), but the vd of NH3 is approximately ten times higher than the vd of NO2, which leads to a higher contribution of NH3–N to the total deposition, particularly for conifer forests with their year-round present foliage (Figs. 2a and 3). The dry deposition of NH3 was especially large at sites close to intensive agriculture (42–65%, median 59% of the total N deposition). Both the NH3 concentration and the NH3–N deposition are significantly higher with more lifestock units (p < 0.01 for both). They were highest on the Plateau (median: 12.4 kg ha−1 a−1) and lower in the other regions (median annual deposition in kg ha−1 SA: 4.6, CA: 0.6, NA: 1.9, J: 4.8). NO2 concentrations and NO2–N deposition were also highest on the Plateau (median annual deposition in kg ha−1 SA: 1.4, CA: 0.3, NA: 0.5, P: 2.6, J: 1) and the depositon was significantly higher with higher number of inhabitants within 100 m and 1000 m (p < 0.05 for both). The N compounds HNO3, aerosol-NH4+, and aerosol-NO3- only accounted for 4–18% (median 10%) of the total inorganic N deposition (see Table S1). The concentration of gaseous HNO3 was very low, yielding low deposition of this compound. The aerosols showed slightly higher concentrations in air, however they have very low vd and thus a low deposition rate. N deposition of the three compounds was highest in the Plateau (median: 2.3 kg ha−1 a−1) and lowest in the Central Alps (median: 0.6 kg ha−1 a−1), with similar values in the Northern Alps (median: 1.8 kg ha−1 a−1), the Southern Alps (median: 1.8 kg ha−1 a−1), and the Jura (median: 1.6 kg ha−1 a−1).

An analysis of variance (ANOVA) was used to test for the difference in the total inorganic N deposition and individual N compound between regions, land use types, and ecosystem types. A linear regression analysis was used to test for changes in measured N deposition against time. Annual decrease was calculated based on the slopes gained by the regression analysis. All analyses were performed in R 3.0 (http://rproject.org). 3. Results 3.1. Regional variation of inorganic N deposition The total annual inorganic N deposition of the year 2014 differed significantly between the regions, with deposition values ranging from 4 to 53 kg N ha−1 a−1 (median annual sum of each region in kg ha−1 SA: 26.7, CA: 4.7, NA: 14.5, P: 26.5, J: 14.6; p < 0.001; see Table 1 for abbreviations). The highest total N deposition rates were found in the Southern Alps (MAG, SARE, NOV), and the lowest deposition values were found at higher elevation in the Central Alps. The total inorganic N deposition was significantly related to the surrounding land use type (p < 0.001) and was highest for intensive agriculture (median: 48.7 kg N ha−1 a−1) and meadows within intensive agriculture (median: 32.4 kg N ha−1 a−1). It was lower for meadows, pastures and non-agricultural land use areas, although there was still considerable variation between the sites (land use surrounding site, see Table 1). In the open ecosystem types (alpine/subalpine grasslands and mountain hay meadows) the N load via precipitation was the main contributor (median 71% and 62%, respectively), whereas in the forest ecosystems the dry deposition contributed the most (median 70%), with NH3 being the major contributing compound (23–65%, median 42%). At most sites, NO2, HNO3, and the aerosols NH4+ and NO3− contributed only little to the total N deposition (7–37%, median 17%). The contribution of reduced N deposition (41–85%, median 63%) is mostly higher than of oxidised N deposition.

3.2. Comparison to CLN The comparison between the measured inorganic N deposition values and the CLN values for each ecosystem (Table 1, Fig. 3) revealed that at 29 out of 34 sites, measured inorganic N deposition values fell above the lower limit of the CLN range. Only at high elevation sites in the Central Alps (ARD02, FUR, NAT, CEL, DAV) the lower limit of the CLN was not exceeded; however, at most of these sites, the N deposition was still very close to the lower CLN limit. At 20 sites, the N deposition surpassed the upper limit of the CLN range, even by a factor of more than two at five sites. At seven sites, the N deposition was 1.5–2 times higher than the upper limit of the CLN range. The exceedance of the CLN thresholds varied between different ecosystem types, as did the contributions of the N components. For alpine/subalpine grasslands, the upper limit was only reached or exceeded at three sites, while the lower limit was exceeded at seven sites. The main contributor to total N in grasslands was deposition by precipitation (55–85%, median 71%) and NH3 only added little to the exceedance at most sites (7–32%, median 11%). Conversely, most forest and bog sites revealed a strong exceedance of the lower limit (bogs: nine sites; forests: all sites) and the upper limit (bogs: eight sites; forests: nine sites) as well as a stronger influence of NH3 (bogs: 12–64%, median 36%; forests: 23–65%, median 42%). The data also shows that at 22 sites, the lower limit of the CLN is already exceeded with the N deposition from precipitation alone.

3.1.1. Inorganic nitrogen deposition in bulk The N deposition by precipitation (bulk deposition) was the largest contributor to the total inorganic N load at 16 sites (Fig. 2a), including mainly open ecosystems such as alpine/subalpine grassland, mountain hay meadows, and raised bogs, but not forests and bogs. NH4–N contributed 4.5–50% (median 26%) to total N deposition and NH4–N weighed in slightly more (40–74%, median 57%) to bulk deposition than NO3–N. NH4+ in precipitation was more strongly influenced by local or regional sources than oxidised N compounds. The NH4+ concentrations in the bulk deposition samples (Table 2) were high in the Southern Alps, low in the Central Alps, and mid-range in other regions of Switzerland. The same pattern is found in the NH4–N deposition (median annual deposition in kg ha−1 SA: 10.3, CA: 1.4, NA: 5.9, P: 4.6, J: 3.3; Fig. 3). NO3–N deposition added 6–63% (median 19%) to the total inorganic N deposition. NO3–N deposition depended both on the NO3− concentration in precipitation (Table 2) and on the amount of precipitation. The Southern Alps had high NO3− concentration as well as high precipitation, leading to high NO3–N deposition (median: 7.8 kg ha−1 a−1). Sites on the Plateau showed only slightly lower NO3− concentrations than those in the Southern Alps, but the amount of precipitation was relatively low, thus the NO3–N deposition by precipitation was lower than in the Southern Alps (median: 3.1 kg ha−1 a−1). NO3–N deposition by precipitation was lowest in the Central Alps (lower NO3− concentrations and low amount of precipitation, median: 1.6 kg ha−1 a−1). The concentration in the Northern Alps and the Jura

3.3. Trends over time The development of the N deposition was observed at five sites over a period of 16–18 years (2000/2002–2017; Fig. 4; Table S10). A significant decrease in oxidised compounds was found at four out of five sites (CHA: 1.7% per year, p < 0.001; MAG: 1.8% per year, p < 0.001; RIG: 1.6% per year, p < 0.001; SCH: 1.8% per year, 219

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Fig. 2. a) Portion (in %) of the different N compounds per study site. The size of the circles is proportional to the square root of the nitrogen deposition at each site. b) Illustration of N emission sources and the precipitation at each site. The bars do not represent absolute numbers but are adapted for illustration purposes. The number of inhabitants per ha derives from data provided by the Swiss Federal Office of Statistics. The livestock units are recorded as part of the sampling site description. The precipitation amounts equal the precipitation indicated in Table 1.

p < 0.01; MAG: p < 0.001; RIG: p < 0.01, Fig. 4f). The bulk deposition decreased significantly at all sites (BA: decrease of −1.4% per year, p < 0.01; CHA: 2% per year, p < 0.01; MAG: 1.5% per year, p < 0.05; RIG: 1.2% per year, p < 0.05; SCH: 2.1% per year, p < 0.001; Fig. 4c). Dry deposition however only changed significantly at MAG with an increase of +1.2% per year (p < 0.01; Fig. 4d). The contribution of bulk versus dry compounds changed significantly at two sites with dry deposition becoming more important (CHA: p < 0.05; MAG: p < 0.01; Fig. 4e). The amount of precipitation, while variable, did not change significantly over time.

p < 0.001; BA: n.s.; Fig. 4a). In contrast, the reduced N compounds declined significantly at only two sites (SCH: 1.4% per year, p < 0.05 and BA: 1% per year, p < 0.05), while at one site, these compounds increased significantly (MAG: +1.2% per year, p < 0.05, Fig. 4b). Nevertheless, the total N deposition still surpassed the CLN at SCH and BA in the year 2016 and 2017, respectively. Overall, the amount deposited, as well as the year-to-year variability in deposition, was larger for reduced N compounds than for the oxidised compounds. The contribution of oxidised versus reduced compounds changed significantly at three sites with reduced compounds becoming more important (CHA:

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Central Alps, showed higher NH4+ concentrations than the other four Central Alps sites. The site is within 200 m of several farms and therefore likely affected by the N emission of these farms. NH3 mainly originates from agriculture (Buijsman et al., 1987; ECETOC, 1994) and is mostly deposited close to the source, thus NH3 concentration and deposition is especially high close to agricultural activity. High livestock densities, intense agricultural activities, excessive manure management (incl. storage), and the application of fertilizer, may increase ammonia emissions and therefore increase the fraction of the total inorganic N deposition that is ammonia (Huang et al., 2007; Hutchings et al., 1996). Therefore, NH3 contributes strongly to N deposition on the Plateau, where the agricultural intensity is high. Due to the large surface area of forest canopies, the vd and therefore the deposition of NH3 into forests, is especially high. Thus, at eight forest sites close to agricultural activities, the N deposition caused by NH3 was substantial (30–65% of the total N deposition). Direct measurements in canopies of deciduous and conifer trees revealed higher NH3 uptake rates in summer-green leaves than in evergreen needles (Adriaenssens et al., 2012; Gessler et al., 2003, 2000). Stomata conductance, thus stomata aperture did not fully explain NH3 fluxes into the foliage. It has been assumed that NH3 deposited on plant, especially on leaf surfaces was also metabolised by the plants (Gessler et al., 2003). The here used vd for NH3 of 30 mm s−1 for conifer forests accounted for the year-round presence of needles and the high canopy roughness. For deciduous and mixed forests, smaller vd for NH3 was applied. Different grass species differed in leaf conductance and thus in NH3 plant uptake, but different stomatal conductance at leaf level were largely affected and partly offset by the total number and leaf area per plant (Hanstein et al., 1999). It also has been emphasised that leaf age strongly influences NH3 uptake in graminoids (Hill et al., 2002). Re-emission of NH3 from plants and soils to the atmosphere in croplands and fertilised areas depends largely on the fertilizer amount, the timing of the fertilizer application and the climatic conditions (Krupa, 2003; Sutton et al., 2013). For semi-natural grassland without direct fertilizer applications, re-emission of NH3 from plant to the atmosphere is assumed to be negligible, resulting in a net plant uptake of atmospheric NH3. Vd for NH3 in croplands is therefore lower than for semi-natural grassland, accounting for the re-emission of NH3 under fertilizer addition. Gaseous NO2 generally contributed little to the total N deposition and may only have played a role at sites close to intense traffic (Hueglin et al., 2006). For instance, dry deposition of NO2 accounted for 23% and 24% of the total N deposition at VIS and OTH, respectively. Both sites are close to major roads. Generally, NO2–N deposition is related to number of inhabitants surrounding the site and was highest in the Plateau with the highest population density and traffic volume. The remaining N compounds HNO3, aerosol-NH4+, and aerosolNO3- also contributed little to the total N deposition in our study. Even at sites with little precipitation and/or far from agricultural activities (VIS, DAV, BET, NAT, JUS) these three compounds weighed only around 15% to the total inorganic N deposition. In general, they may play a significant role in more arid climates than the studied here (e.g. Galy-Lacaux and Delon, 2014). Even though, some literature suggests that vd for HNO3 can vary (e.g. Walcek et al., 1986) we used a constant value for vd of HNO3 because HNO3 is rapidly deposited irrespective of the plant canopy resistance (e.g. Duyzer and Fowler, 1994). Even if employing the same vd adjustment based on ecosystem for HNO3 as those for NH3, we estimated an increase of total inorganic N deposition by 0–5% only, thus an ecosystem specific vd for HNO3 would not alter the results presented here. The total inorganic N load at each site correlates with the presence of emission sources (agricultural activities and population density) and the influence of precipitation illustrated in Fig. 2. At WAU, however, the high inorganic N load does not seem to correlate with these emission sources. Nonetheless, the high loads – especially of ammonia – can

Table 2 Weighted mean annual NH4−-N and NO3 --N concentrations in precipitation, gaseous NH3–N, NO2 and HNO3–N and aerosol concentration for NH4+-N and NO3−-N all 34 sites.

4. Discussion 4.1. Nitrogen deposition The differences in the total annual inorganic N deposition between the five Swiss biogeographic regions were mainly driven by varying amounts of precipitation, differences in concentrations of NO3− and NH4+ in precipitation as well as by dry deposited NH3 (Table 2). At sites far away from intensive agriculture and roads (no NH3 and NO2 sources), the wet (bulk) deposition dominated the total N deposition. Therefore, the regional N deposition depended strongly on the amount of precipitation in addition to the concentration of NO3− and NH4+. The highest total N deposition rates were found in the Southern Alps because of the high precipitation levels caused by orographic lift (Rogora et al., 2016), combined with emissions from the Po Valley. Sites in the Central Alps that are farther away from N sources and screened from precipitation by high mountains on both sides, had the lowest N deposition rates. In addition to regional conditions, atmospheric N input to the sites was also influenced by local sources. The total N deposition depended strongly on the surrounding land use type and was considerably higher at sites close to intensive agriculture. This highlights the importance of agricultural practices and the proximity to agriculture for N emissions as well as for local N deposition in Switzerland (EKL, 2014). For instance, the site ARD 06, although situated within a meadow in the 221

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Fig. 3. The deposition of the seven N components in (kg N ha−1 a−1) at the 34 sites in the year 2014, sorted into different ecosystem types and regions: a) alpine/ subalpine grassland, b) mixed forests, c) coniferous forests, d) deciduous forests, e) mountain hay meadows and bogs, and f) raised bogs. See Table 1 for abbreviations of the regions.

high depositions at these sites are most likely caused by high precipitation, the intensive agricultural activities in the surrounding of the sites, or the proximity to a major road. This is similar for sites where the N deposition was 1.5–2 times higher than the upper limit of the CLN range. Unfortunately, high exceedance of the CLN thresholds were observed for some bogs (MAG, WAU) and raised bogs (HUD), although these ecosystems represent fully protected areas and are often inhabited by endangered (red-listed) plant and animal species. The high input of N into these areas most likely contributes to the observed decline in the quality of Swiss bog vegetation (Feldmeyer-Christe and Küchler, 2017). Protection of these sensitive ecosystems may only be possible by enlarging the protected areas or adding buffer zones with clearly restricted agricultural activities. High exceedance of the CLN thresholds was also observed for forest sites. High N input into forests leads to acidification of the forest soil, nutrient imbalances, disturbance in fine root formation, changes in

be explained: WAU is situated in the area with a very high cattle density (Platform of Swiss Academy of Sciences (SCNAT), 2018), but the livestock units are ascribed to the locations of the farm houses instead of the livestock housing and grazing lands, or the area of manure application. Thus, although WAU is surrounded by agricultural land where manure is intensively applied (leading to high NH3 input), the farm houses are not nearby and so the livestock units are not ascribed to the area of the sampling site. 4.2. Comparison to CLN The inorganic N deposition exceeded the lower limit of the CLN range at most sites and at many sites, even surpassed the upper limit of the CLN range. This highlights that the N deposition is still too high for the most sensitive ecosystems in Switzerland. The highest exceedance of the CLN range was at NOV (alpine/subalpine grassland), MAG (bog), HUD (raised bog), WAU (bog), VIS (forest) and SCH (forest) sites. The 222

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Fig. 4. The temporal trends of the total inorganic N deposition at five sites for a) oxidised b) reduced N compounds and the change in c) bulk and d) dry N deposition. e) shows the change in contribution of bulk and dry f) the change in the contribution of oxidised and reduced deposition.

implementation of catalytic converters in cars and DeNOx systems in industrial settings from 1980 onwards. Whereas reduced compounds decreased slightly but significantly at two sites only and showed significant increase at one site (which may indicate the presence of a new NH3 source) no change at two sites despite abatement strategies to reduce N emission in the agricultural sector. This emphasises that the measures taken to reduce NH4+ and NH3 emissions have not been sufficient so far to effectively reduce the N input into sensitive ecosystems. The measures to reduce N emission from agriculture in Switzerland are non-binding recommendations, and they therefore may also not have been applied consistently. Additionally, there has been further intensification of agriculture and changes in agricultural practices (e.g. open stables due to animal rights, import of N-rich fodder), which may increase overall N emission and offset any reduction measures taken.

mycorrhizae, and increased sensitivity to secondary stresses (de Witte et al., 2017; Erisman and de Vries, 2000; Krupa, 2003) and susceptibility to storm damage (Braun et al., 2003; de Witte et al., 2017; Mayer et al., 2005).

4.3. Trends over time The trends of the inorganic N deposition from 2000/2002–2017 are in line with the reported emission data for Switzerland (FOEN, 2018). From 2000 to 2016, emissions of NOx decreased strongly (−2.4% per year, p < 0.001), but emissions of NH3 decreased only a very slightly (−0.4% per year, p < 0.001). These decreases are also apparent in Europe-wide trends (Colette et al., 2016). The significant decrease in oxidised compounds at four out of five sites may be due to the 223

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The bulk deposition decreased over time at all sites and at some sites also the contribution of bulk deposition to total N deposition decreased, despite there being no changes in the amount of precipitation. Dry deposition did not change. It is possible that, due to the increase in average temperature, less gas is dissolved in precipitation.

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5. Conclusions Our study highlights that – while there are regional differences in N deposition caused by differences in population density, traffic patterns, and the degree of agricultural intensity – most sensitive ecosystems in Switzerland are still exposed to excessive inorganic N loads by atmospheric deposition, with negative consequences for the biodiversity and stability of these ecosystems. In Switzerland, forests for example are not only an important habitat for many plant and animals species, they also play an essential role in the protection against hazards such as erosion, avalanches, and landslides (Brang et al., 2008; Olschewski et al., 2012). Also, the semi-natural subalpine and alpine grasslands are rich in biodiversity, and soils in steep slopes are only secured through a speciesrich vegetation. While measures to reduce oxidised N compounds seem to have been partially successful, reduced compounds – especially gaseous NH3 – show little to no decrease over time. In Switzerland, these measures are currently only recommendations. Additional measures aiming at reducing the N input into sensitive ecosystems are urgently needed and should be legally enforced. In particular, measures to decrease the emissions of reduced N compounds in the agricultural sector should be put into action immediately. Measures could be linked to national direct payments in order to accelerate the implementation processes. In order to protect all sensitive ecosystems, and to guarantee the survival of often rare species living in these ecosystems, we strongly recommend enlarging the protected areas and to set-up buffer zones around these areas. Especially for bogs, limited land use should be aimed for preserving the life conditions for rare species (e.g. removal of biomass late in growing season, provision of an adequate water level). Declaration of interests The authors declare that they have no known competing financial interests or personal relationships that could have appeared to influence the work reported in this paper. The authors declare the following financial interests/personal relationships which may be considered as potential competing interests. Acknowledgements We would like to thank to the Swiss Federal Office for the Environment (FOEN), and the cantons of GR, AG, BE, GE, SO, TI and OSTLUFT for financial support of this study. We acknowledge The Swiss Federal Laboratories for Materials Science and Technology (Empa) for providing the data for N in precipitation and NO2 for the NABEL stations, and Patrick Schleppi for the bulk precipitation data of the ANA B site. We thank Beat Rhim (MeteoTest) for his continuous and supportive advice. Our thanks go to all the field and laboratory assistants for changing the samplers and analysing the samples, and to the leading project managers of the various institutions. Appendix A. Supplementary data Supplementary data to this article can be found online at https:// doi.org/10.1016/j.atmosenv.2019.05.005. References Aber, J.D., Magill, A., Mcnulty, S.G., Boone, R.D., Nadelhoffer, K.J., Downs, M., Hallett, R., 1995. Forest biogeochemistry and primary production altered by nitrogen

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