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Elevated ozone and nitrogen deposition affect nitrogen pools of subalpine grassland
Environmental Pollution 201 (2015) 67e74
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Elevated ozone and nitrogen deposition affect nitrogen pools of subalpine grassland €ch, Alain Valsangiacomo, Jochen Mayer, Hans-Rudolf Oberholzer, Seraina Bassin*, David Ka Matthias Volk, Jürg Fuhrer Agroscope, Institute of Sustainability Sciences ISS, Reckenholzstrasse 191, CH-8046 Zurich, Switzerland
a r t i c l e i n f o
a b s t r a c t
Article history: Received 5 December 2014 Received in revised form 25 February 2015 Accepted 27 February 2015 Available online
In a free-air fumigation experiment with subalpine grassland, we studied long-term effects of elevated ozone (O3) and nitrogen (N) deposition on ecosystem N pools and on the fate of anthropogenic N. At three times during the seventh year of exposure, N pools and recovery of a stable isotope tracer (15N) were determined in above- and belowground plant parts, and in the soil. Plants were much better competitors for 15N than soil microorganisms. Plant N pools increased by 30e40% after N addition, while soil pools remained unaffected, suggesting that most of the extra N was taken up and stored in plant biomass, thus preventing the ecosystem from acquiring characteristics of eutrophication. Elevated O3 caused an increase of N in microbial biomass and in stabilized soil N, probably resulting from increased litter input and lower litter quality. Different from individual effects, the interaction between the pollutants remained partly unexplained. © 2015 Published by Elsevier Ltd.
Keywords: Ozone Nitrogen Grassland Stable isotopes Soil
1. Introduction Concentrations of reactive nitrogen compounds (N) and tropospheric ozone (O3) have risen substantially since the industrial revolution (Galloway et al., 2004) and are expected to increase furthermore in the future (Ashmore, 2005). Both trace gases are known to threaten the biodiversity and functioning of ecosystems separately (Sutton et al., 2014), but large uncertainty exists about interactive effects between the two pollutants on (semi-)natural vegetation. Anthropogenic N input has been shown to affect plant productivity, community composition, and biodiversity especially in nutrient-poor semi-natural habitats (Bobbink et al., 2010). The impact of anthropogenic N on ecosystem functioning is modulated by the amount of N retained in soil and vegetation pools, whether it is stored, released to microorganisms and plants, or lost from soil via gaseous fluxes and leaching. In a comprehensive review including different temperate (semi-)natural ecosystems Phoenix et al. (2012) revealed both mineralization and nitrification to increase in five out of ten cases under elevated N deposition.
* Corresponding author. Agroscope, Climate/Air Pollution enholzstrasse 191, CH-8046 Zurich, Switzerland. E-mail address: [email protected] (S. Bassin). http://dx.doi.org/10.1016/j.envpol.2015.02.038 0269-7491/© 2015 Published by Elsevier Ltd.
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Artic and alpine ecosystems are considered especially vulnerable to increasing N deposition since they have been exposed to very low deposition levels in the past. These soils are still nutrient poor and rich in organic matter due to the temperature limitation of € rner, 2003). As a consequence, the soil solution is mineralization (Ko dominated by dissolved organic N (DON) and traces of NHþ 4 , while free NO is typically missing (Chapin et al., 2012). In spring after 3 snowmelt, an N pulse is released from the decaying winter microbial biomass, providing a DON source for plants and microbes during the growing season (Schmidt et al., 2007). When mineral N is deposited into nutrient limited ecosystems, it is often instantly immobilized by soil microorganisms, since these are mostly C and N co-limited, especially in arctic and alpine soils (Chapin et al., 2012). €rner, With plants in these habitats being primarily N limited (Ko 2003), and in view of their capacity to take up and store excess N (Monson et al., 2006) the plant-microbe competition for N is €rner, 2003). Thus, some of the anthroconsidered to be severe (Ko pogenic N will accumulate in soil microorganisms and plants, reducing the risk for mineral N leaching for a certain time (Phoenix et al., 2012). However, remarkable differences exist among habitat types concerning the ecosystem pools that act as storage sites of the pollutant N: in lowland heath, two-thirds was retained in microbial biomass while in grasslands, bulk soil and plant pools were the dominant sinks (Phoenix et al., 2012). The stable isotope technique using a 15N enriched tracer is a widely used method to follow the
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fate of anthropogenic N in environmental systems (Dawson et al., 2002). In a Swiss mountain meadow, a substantial part of added tracer was instantly immobilized in soil, presumably bound to clay particles (Providoli et al., 2006), in line with results from a worldwide review on 15N tracer recoveries in temperate and arctic ecosystems (Templer et al., 2012). O3 is the most phytotoxic air pollutant in Europe (Ashmore, 2005). It primarily affects the carbon (C) budget of plants by reducing CO2 assimilation and by increasing the C investment into chemical and physical defense mechanisms (Andersen, 2003). But it has been shown to affect N budgets as well, e.g. by impairing N resorption from leaves before abscission, causing a substantial increase in whole-tree N loss through litter fall (Uddling et al., 2006). O3-induced changes in litter chemistry, which often include increased levels of phenolic compounds (Wang and Frei, 2011) may lead to reduced decomposition rates as shown in hardwood forests (Liu et al., 2009). In soybean, changes in litter decomposition processes were identified as cause for the increased levels of total N found in soil after six years of elevated O3 exposure (Pereira et al., 2011). Often, microbial activity was reduced by O3 fumigation as a result of reduced C allocation belowground (Andersen, 2003). While the direct impact of O3 on N budgets are expected to be small compared to those of increased N input, the magnitude of the indirect effects through reduced C availability is unknown. The interactive effects of elevated O3 and N deposition have been studied mainly in pot experiments with a focus on biomass production (e.g. Wyness et al., 2011) revealing N to alleviate the negative effects of O3 in most cases. Yet, long-term studies on N pools are missing in this context. Using a free air fumigation system we were able to expose intact monoliths of a species-rich subalpine grassland in situ at 2000 m a.s.l. to different combinations of O3 and N deposition for seven years. Substantial increases in aboveground biomass were recorded in response to N addition, mainly as a result of an extraordinary success of sedges, which tripled their relative abundance up to 30% (Bassin et al., 2013). In contrast, the O3 treatment did not affect above-and belowground biomass production (Volk et al., 2014) but it accelerated leaf senescence (Bassin et al., 2007b). The aim of the present study was to investigate the long-term effects of elevated O3 and N deposition, singly and in combination, on the soil and plant N pools of the subalpine pasture and the fate of anthropogenic N in this ecosystem. After a pulse application of a dual labelled 15NO15 3 NH4 tracer in May, soil cores were taken at three dates during the seventh and last growing season of the experiment and the ecosystem N pools (green phytomass, necromass, litter, roots, microbial biomass, soluble soil N, extractable soil N, stabilized soil N), their C:N ratio, as well as the recovery rates of the tracer was determined to test the following hypotheses: 1) The major fraction of added mineral N is bound to the mineral and organic fraction of the soil. 2) Elevated O3 causes accumulation of soil N in the form of soil organic matter (SOM), as a result of reduced microbial activity (due to lesser availability of labile C) in combination with less degradable litter. 3) Consequently, highest N accumulation in soil pools is expected when elevated O3 and N are combined.
2. Methods 2.1. Experiment The experiment was established at Alp Flix, a high plateau at 2000 m a.s.l. near Sur, Grisons, Switzerland (46 310 N/9 380 E). The
climate is characterized by a mean annual temperature of 1.1 C, a vegetation period lasting from April to October with average temperatures of 6.2 C and a permanent snow cover in winter (Volk et al., 2014). The annual sum of rainfall is approximately 1280 mm. The vegetation under investigation, a Geo-MontaniNardetum, is a typical perennial alpine pasture community covering large areas of the subalpine zone in the Alps and the Pyrenees. Grasses like Festuca rubra SCHLEICH. ex GAUD. and Nardus stricta L. and sedges like Carex sempervirens VILL. are the most abundant species (Bassin et al., 2013). The soil is a slightly acidic Cambisol with a depth varying between 20 and 40 cm, developed on a Serpentine bedrock. Rooting depth is 10e20 cm, with 80% of the roots found in the uppermost 7 cm. The experiment consisted of nine rings (Ø 7 m) of a free air O3 fumigation system with three different levels: ambient air (ca. 47 ppb); 1.35 ambient concentration, 1.73 ambient concentration, mean enrichment factor for the seven experimental years (for details see Volk et al., 2014). Rings were arranged at random in three blocks. Each ring contained 20 turf monoliths (40 cm long 30 cm wide 20 cm deep) excavated out of a nearby GeoMontani-Nardetum pasture, put into drained plastic boxes and randomly assigned to the rings. Five N levels equivalent to 0, 5, 10, 25, and 50 kg N ha1 y1 were applied in the form of ammonium nitrate dissolved in 200 ml of well water during the growing season every second week (solute concentration: up to 0.018 g N l1). Average wet and dry background deposition amounted to 4 kg N ha1 y1 (Bassin et al., 2007b). For the present study, we used a subsample of 36 monoliths exposed to either the highest concentration of O3 (1.73 ambient; “þO3”), the highest N-deposition (50 kg N ha1y1; ”þN00 ), or the combination of these two (“þN þ O3”), as well as from the untreated control (n ¼ 9 per treatment combination). 2.2. Sampling collection and processing The field work was accomplished in 2010, the seventh and last year of treatment. On 11 May, shortly after snowmelt, 0.04 g m2 of 15 4 N in the form of dual labelled 15NO15 3 NH (60 atom %) dissolved in 2.027 l water was sprayed onto the vegetation in the selected 36 monoliths, followed by the standard N application (see above). On 13 May, 12 July and 30 August, two cores each with a height of 10 cm and a diameter of 6 cm were taken in the monoliths. They were kept cool in boxes and were taken to the laboratory on the same day, where they were stored at 4 C until processing. Soil cores were halved and the halves were merged together to form two mixed samples of which one was sieved (2 mm mesh size) to get a root-free soil sample. Roots were removed with tweezers and finer roots that passed the sieve were recaptured during further processing of the soil (see below). The other sample was used to determine root biomass by washing on a sieve with 0.063 mm mesh size. Aboveground phytomass was cut at the soil surface of each core and green phytomass was separated from standing-dead (subsequently called “necromass”) and leaf litter (“litter”). On 12 July, aboveground biomass was harvested in two portions, i.e. above and below 2 cm, which was the standard cutting height used in the experiment for yield assessment (>2 cm ¼ “green phytomass removed through harvest”). All plant and soil samples were weighed after drying to constant weight at 60 C and then ground to a fine powder with a ball mill (Mixer Mill, Retsch). In the sieved fresh soil, microbial biomass C and N was determined using the chloroform fumigation extraction method (Brookes et al., 1985). Micro-roots were separated from the soil by a pre-extraction step (Olfs and Sherer, 1996). For this purpose 100 g fresh soil was extracted with 240 ml 0.05 M K2SO4 for 30 min. The
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interactions of them was made using likelihood ratio tests (Burnham and Anderson, 2002). Data were transformed when necessary. All analyses were carried out by means of the statistics software R (R Development Core Team, 2011) using the packages nlme, nortest, lattice, and car.
soil water suspension was collected in a beaker and left for 30 min to let the soil settle. The floating roots were sucked with a vacuum pump. Subsequently, the soil water suspension was filtered on a 589 paper filter (Schleicher and Schüll, Germany). The N obtained in this pre-extraction step was considered “soluble soil N”. The filtered soil was then divided into two halves. One halve was immediately extracted with 120 ml 0.5 M K2SO4 for 30 min (“extractable soil N”). The second half was fumigated with 25 ml chloroform in a desiccator for 24 h at 25 C and later extracted with 120 ml 0.5 M K2SO4. Total organic C (TOC) and total N (TNb) in soil extracts were determined by infrared spectrometry after combustion at 850 C (TOC) and subsequently by chemoluminescence (TNb) (DIMATOC 100, Dimatec, Essen, Germany). Microbial biomass C and N was calculated according to Joergensen (1996), using a kEN value of 0.54 (Joergensen and Mueller, 1996) and a kEC value of 0.45 (Joergensen, 1996). The 15N content of the solutions obtained in the three soil extraction steps was determined using a modified diffusion technique described by Mayer et al. (2003). The diffusion filters, the ground plant and soil samples were analysed for N and 15N content using a Flash EA1112 Series elemental analyzer (Thermo Italy Rhodano, Italy) coupled to a Finnigan MAT Delta plusXP isotope mass spectrometer (Finnigan MAT, Bremen, Germany). The 15Nisotope composition were expressed in standard notation (d15N) in parts per thousand (‰) relative to atmospheric N2 (standard). The d15N in the microbial biomass (MB) was calculated according to Mayer et al. (2003). All d 15N values were obtained by correcting the 15N enrichments with the background values of the N pools without labelling (green phytomass 2.8, necromass 2.6, roots 1.8, litter 2.4, soluble soil N 4.5, extractable soilunfum N 8.1, extractable soilfum N 3.1, stabilized soil N 2.7 d 15N). The 15N amount of the fumigated and extracted soil (“stabilized soil N”), soluble soil N (N extracted by 0.05 M K2SO4 solution), extractable N (N extracted by 0.5 M K2SO4 solution), and microbial biomass was calculated by multiplying its total N with the respective 15N abundance.
3. Results 3.1. Ecosystem properties Averaged over the growing season, in the control plots total ecosystem N pool was 364 g m2, of which the major part was stabilized in soil (317 g N m2) (Fig. 1a). The plant pools accounted for 5.8 g m2 aboveground, and 19.3 g m2 belowground. In addition, 18.3 g N m2 was stored in microbial biomass, while only 3.1 g m2 was soluble and extractable from soil. In July, approximately 1.8 g N m2 were allocated from roots to shoots, of which 1.2 ± 0.2 g m2 were removed through harvest (Fig. 1a), but by the end of the season, in August, N pools in living plant parts were almost equally distributed as in spring. N pools in necromass and litter decreased during the growing season by 11% and 65%, respectively (Figs. 1 and 2). 3.2. Tracer recovery Of the added stable isotope tracer, 67% were recovered two days after application in the control plots (Fig. 1b). 31 ± 4% were found in roots, 11 ± 1.1% were stabilized in soil, 12 ± 2% were found in necromass, 4.3 ± 0.7% in green phytomass, and 5 ± 1.4% in microbial biomass. In July, values decreased to 23 ± 4% in roots and increased to 14 ± 0.6% in green phytomass, of which 30% were exported through harvest. By the end of the season, 63% of the added tracer were recovered, 14.8 ± 1.7% were stabilized in soil, 25 ± 0.9% were found in roots, 9.1 ± 0.7% in necromass, 8.7 ± 0.7% in green phytomass, 4.3 ± 0.03% in microbial biomass and 0.6 ± 0.01% in litter. In most of the pools, the tracer recovery decreased significantly over time, while it increased in stabilized soil N (Fig. S1 and Table S2).
2.3. Statistical analyses 3.3. Nitrogen effects Effects of treatments on N pools and C:N ratios were tested in a repeated measures split-plot analysis using a linear mixed effect model (lme) (Pinheiro and Bates, 1996) with O3 at the main plot level (n ¼ 3) and N at the sub-plot level. All parameters were repeated by sampling date. Block, O3, and N entered the model as class variables, whereas time was a continuous variable. To account for the inherent variability among monoliths, initial grass cover (Grass2004), initial N content (Nsoil2003) and pH of the soil measured in 2003 before treatment onset were used as continuous co-variables. Inference on the significance of variables and selected
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Extra N addition increased the following N pools significantly compared to control plots (Fig. 2, þN and þN þ O3): green phytomass (þ31%) necromass (þ41%), litter (þ30%), roots (þ31%), and soluble soil N (þ22%). The C:N ratio was significantly reduced in green phytomass (12%), necromass (13%), roots (19%), and soluble N (17%) (Fig. 3), thus indicating that the extra N increased both the size as well as the N concentration of these pools by 20e30% and 10e17%, respectively (data not shown). In contrast, N pools and C:N ratio of microbial biomass, extractable soil N, and
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Stabilized soil N Extractable soil N Soluble soil N Microbial biomass Roots Litter Necromass Green phytomass Green phytomass removed through harvest
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Fig. 1. a) Ecosystem N pools and b) the recovery of a stable isotope 15N tracer in subalpine grassland monoliths in the top 10 cm of untreated plots (control) at three dates during the growing season 2010 (13 May, 12 July, 30 August). The 15NH15 4 NO3 tracer was applied right after snowmelt on 11th May. At harvest, green phytomass above 2 cm was removed.
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Fig. 2. N pools of green phytomass, necromass, litter, roots, soluble N, extractable N, microbial biomass, and N stabilized in soil in plots exposed to elevated O3 concentrations (þO3), elevated N deposition (þN), the combination of them (þN þ O3), and in the control. Means ± 1SE (n ¼ 9) are displayed for three dates during the growing season 2010. Closed circles: control; open circles: þN; closed triangles: þO3; open triangles: þN þ O3. Significant factors and covariates (P < 0.05) inferred from likelihood ratio tests using linear mixed effect models are listed in the figures (for details see Table S1).
stabilized soil N remained unaffected. Only soluble N was increased by 22% throughout the growing season. N removed though harvest in þN and þN þ O3 treatments (2.6 ± 0.3 g m2) was twice the amount in the control (1.2 ± 0.2 g m2). However, below the cutting height, green phytomass still accounted for 2.47 ± 0.4 g N m2, but the N effect was only þ20%. Stable isotope recovery in N pools did not differ between Ntreated and control plots (Table S2), with three exceptions: compared to the control plots (5 ± 1.3%), exported green phytomass contained significantly more 15N (9 ± 1.6%) in N-treated plots (þN and þN þ O3), while 18% less 15N was stabilized in soil (Fig. S1). In the same monoliths, more tracer was recovered in soluble and extractable N, as well as in microbial biomass two days after tracer application compared to control plots, with the effect decreasing over time (significant N time interaction). 3.4. O3 effects Elevated O3 increased the N pool in green phytomass by 9.4% (¼
average of þO3 and þN þ O3 for the growing season, Fig. 2) compared to non-fumigated plots (þN and control) through a rise in N concentration (Fig. 3). Also, microbial biomass N pool was significantly increased by 9%, but at a constant C:N ratio (approximately 6.8), indicating that microbial biomass C was simultaneously stimulated by elevated O3. Remarkably, in response to elevated O3, N stabilized in soil increased significantly by 9% from 322 ± 7 g m2 in non-fumigated plots to 352 ± 3 g m2 (¼ average of þO3 and þN þ O3), but with no change in C:N ratio. Elevated O3 fumigation had no significant effect on 15N recovery in any of the ecosystem N pools (Table S2). 3.5. O3 N interactive effects Significant and consistent interactive O3 N effects were found in belowground N pools, namely in soluble soil N, and extractable soil N, with the effect of O3 being consistently negative in nonfertilized plots (þO3), but positive in fertilized treatments (þN þ O3, Fig. 2), compared to the control. For example, soluble soil
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Fig. 3. C:N ratio of green phytomass, necromass, litter, roots, soluble N, extractable N, microbial biomass, and N stabilized in soil in plots exposed to elevated O3 concentrations (þO3), elevated N deposition (þN), the combination of them (þN þ O3), and in control plots. Means ± 1SE (n ¼ 9) are displayed for three dates during the growing season 2010. Closed circles: control; open circles: þN; closed triangles: þO3; open triangles: þN þ O3. Significant factors and covariates (P < 0.05) inferred from likelihood ratio tests using linear mixed effect models are listed in the figures (for details see Table S3). Missing data for litter C content due to the low amount of litter available.
pools were 33% higher in þN þ O3 but 12% lower in þO3 compared to the control. Interactive effects of O3 and N were even more pronounced in the C:N ratio of roots (Fig. 3, Table S3), with values being 9% lower in þN, 9% higher in þO3, but 15% lower in þN þ O3 compared to the control, when averaged over the growing season. In microbial biomass C:N, the response pattern was similar and statistically significant (Fig. 3). 4. Discussion 4.1. Nitrogen effects N pools in plants, microbial biomass, soluble, and extractable N totalled 47.1 g N m2, similar to values for subalpine meadows (Robson et al., 2010) and alpine tundra (Fisk et al., 1998). Relative to this total pool size, the cumulative load of extra N (þ35 g m2 ¼ 7 5 g m2) applied during seven years in the þN treatment was very high, and sufficient to affect both the nutrient balance as well as ecosystem processes. The additional N was
primarily stored in plant pools, which gained not only in size due to increased biomass production, but also in N content through luxury uptake (Fig. 3) (Bassin et al., 2009), thus contrasting our first hypothesis. N capture for storage is common in alpine plants (Monson et al., 2006), as root biomass acts as a long-term reservoir for C and N (Van der Krift and Berendse, 2002). Here, much more N (up to 24 g N m2) was stored in roots compared to the maximum of 5.5 g N m2 found in green biomass and necromass, respectively, at peak vegetation development (Fig. 1). Also the yearly extra export of 1.4 g N m2 through harvest in þN plots is small, which limits the effectiveness of management options aiming to mitigate the impact of anthropogenic N deposition. Necromass played an important role as well, firstly due to its capacity to adsorb and store extra N (Fig. 1b, Fig. S1), and secondly because it showed the most pronounced enhancement in N pool (þ40%, Fig. 2) due to an accumulation of N-rich senescent leaves, related to the increase of sedges such as C. sempervirens (Bassin et al., 2013). Tough fibrous leaves of sedges are characterized by low decomposability (Aerts and deCaluwe, 1997), which is not improved through higher N contents after N addition (Arnone and Hirschel, 1997). Moreover, in
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tussocks of C. sempervirens decomposition rates are lowered by a drier microclimate (Yu et al., 2011). An increase in the abundance of Cyperaceae species in alpine vegetation is often observed in response to elevated N deposition (Bassin et al., 2012; Soudzilovskaia and Onipchenko, 2005). With the prevalence of poorly degrading graminoids, added nutrients are long-term stabilized in necromass, thus preventing the community from quickly acquiring features such as high decomposition rates and high nutrient availability typical of eutrophication (Soudzilovskaia et al., 2007). Plants were much better competitors for added N than microbes, as 48 h after the application of the tracer 35% of 15N was recovered in living plant parts while only 5% were found in microbial biomass. This contrasts results from studies revealing microorganisms as superior competitors for added mineral N (see the review of Kuzyakov and Xu (2013)). This discrepancy could be caused by the fact that in competition studies the tracer is typically injected directly into the rhizosphere, while here, in order to simulate N deposition from the atmosphere, the tracer was sprayed to the top of the canopy, which could have influenced the spatiotemporal pattern of N uptake. However in alpine tundra, plants were also the dominant N consumers, especially in spring when microbial biomass decreased after snowmelt as a result of C limitation (Lipson et al., 1999). It is likely that plants, regardless of their €sholm et al., 2009), prefer mineral capacity to take up organic N (Na N if available, while microbes preferably make use of excess DON in soil solution (Chapin et al., 2012). As a consequence, microbial biomass N pool and C:N ratio remained unaffected by N addition. Similarly, Manning et al. (2008) found no significant effect of N addition on microbial biomass, thus supporting the assumption that N limitation is not widespread among microbes (Treseder, 2008). The low C:N ratio (6.6:1 in control plots) in comparison to usually observed ratios of around 10:1 (Chapin et al., 2012) may explain the absence of N shortage in the present microbial community. Finally, despite the high N-sensitivity of this vegetation in terms of species composition and biodiversity (Bassin et al., 2013), the risk of N-induced acceleration of N cycling, enhanced N leaching or gaseous N losses to the atmosphere seems to be comparatively low. E.g. leachate collected during the growing season 2010 in a subsample of 20 monoliths revealed an accumulated N loss of only 0.15 g m2 (90% as DON, unpublished data), irrespective of N treatment. Evidently the N cycle in this ecosystem is still tightly closed even after seven years of extra N input. But it remains possible that N stored in plants may become mobilized in the long run due to external triggers such as climate warming or altered precipitation patterns, with unknown consequences for speciescomposition, N- and C fluxes, and N-leaching (Phoenix et al., 2012). 4.2. O3 effects In response to elevated O3, the N pool in green phytomass increased by approximately 9%, as compared to non-fumigated plots, due to higher leaf N concentration. In agreement, in a meta-analysis Wittig et al. (2009) found consistently increased N concentrations in O3 exposed tree leaves and explained this by retranslocation of N from prematurely senescing parts of the crown. Alternatively, N resorption from senescing leaves could also be reduced (Uddling et al., 2006) as a result of both impaired phloem loading and apoptosis-like cell death preventing complete degradation of organelles (Fukuda and Greenberg, 2000). Theoretically, the preferential shift in N allocation towards aboveground phytomass could negatively affect belowground N pools, but due to the low shoot:root ratio typical for these subalpine grasslands, the aboveground increase by 0.5 g N m2 would not lead to a detectable
change in the root N pool, which accounts for 14e24 g N m2 at harvest date. Equally, export of N-enriched phytomass through harvest would not cause noticeable changes in the ecosystem N budget. Similar to green phytomass, elevated O3 also increased the N pool in soil microbial biomass throughout the growing season. Similar results were found in O3-fumigated forests and boreal peat land, presumably as a result of altered litter quality (Morsky et al., 2008; Scagel and Andersen, 1997). Here, at elevated O3 a higher proportion of necromass was found in harvested phytomass in several consecutive years (Bassin et al., 2013), and presumably, soft leaves typical of O3-sensitive species (Bassin et al., 2007a) senesce earlier and are rapidly decomposed thus providing a readily available substrate for microorganisms. Preliminary results from the present experiment indicate higher amounts of soil microbial DNA in the O3-treated plots, however only in July, thus supporting the notion of increased substrate availability at the peak of vegetation development (Schneider, pers. communication). But, with root biomass being two to three times larger than aboveground plant € rner, 2003), roots biomass in alpine and arctic ecosystem (Ko remain the dominant C and nutrient input (Smith et al., 2014), and here the C:N ratio of microorganisms mirrored the ratio in response to the treatment combinations (Fig. 3). Thus, increased root turnover rates could be a more likely explanation for the O3-induced stimulation of microbial biomass. While many studies revealed biomass of roots to decrease under O3 fumigation, namely in fast growing establishing plants and crops (Grantz et al., 2006), only a few investigated root turnover rates. These were stimulated in mature O3-stressed trees (Haberer et al., 2007; Pregitzer et al., 2008) as a possible alternative strategy for nutrient acquisition when mycorrhizal infection rates decline (Wang et al., 2014) due to restricted allocation of soluble carbohydrates. However, due to the lack of supporting data, the mechanism leading to the observed increase in the microbial N pool remains open. As a further consequence of elevated O3 exposure, the pool of stabilized soil N increased significantly, in agreement with our second hypothesis, and the C:N ratio remained unchanged (Fig. 3) thus suggesting enhanced accumulation of poorly degradable soil organic matter. Increased root and leaf litter input combined with changed litter quality could be the cause. In soybean, increased levels of total soil N were found after six years of elevated O3 exposure as a consequence of altered litter decomposition processes (Pereira et al., 2011). A low ratio of soluble to structural carbohydrates (Kontunen-Soppela et al., 2007) as well as higher contents of phenolic compounds (Wang and Frei, 2011) are frequently observed physical defence mechanisms against O3 stress, leading to lower degradability of plant material (Booker et al., 2005). In addition, soil C and N turnover rates could have been slowed by the expansion of the slow-growing stress-tolerant N. stricta in þO3 and þN þ O3 plots (Bassin et al., 2013). According to Personeni et al. (2005) the rhizosphere priming effect through root exudates among grass species is dependent on their growth strategy, suggesting lower C and N turnover rates in less competitive species. In the long run, this O3-induced accumulation of SOM has the potential to substantially alter both structure and function of the ecosystem in terms of nutrient cycling, water retention capacity, but also C sequestration rates in a positive way, eventually counterbalancing the negative effects of O3 on C-sequestration expected from growth reductions (see Harmens and Mills, 2012). 4.3. O3 N effects We expected highest N accumulation in þN þ O3 plots as a consequence of reduced litter degradation in þO3 plots in
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combination with increased soil N storage in the þN treatment (third hypothesis). In fact, the largest extractable and soluble soil N pools were observed in this treatment combination (Fig. 2). Also, C:N ratio of roots and microbial biomass was reduced, suggesting highest N availability and quick N turnover rates in þN þ O3. However, the same was not observed for N stabilized in soil, for which the largest pool was found in þO3 (Fig. 2). As an explanation, in þN þ O3 plots N addition could counteract the effect of O3 on SOM accumulation through N-induced accelerated litter decomposition. However, evidence for this explanation remains weak, and because the 15N tracer study did not improve our understanding of the underlying processes (Fig. S1), it cannot be ruled out that besides the co-variables used in the model (soil pH, original soil N content, original grass cover) additional factors were involved in the observed N O3 interaction. Our data suggest that in accordance to aboveground biomass, individual species abundance and performance, the interactive effects of O3 and N on ecosystem N pools were ambiguous. Rather, the two pollutants acted individually: while the effects of N were limited to the plant N pools, O3-induced changes in C metabolism indirectly caused changes in soil C and N urnover processes. But, the interpretation of the results partly relies on assumptions which could not be validated due to a lack of comparable O3 N experiments. 5. Conclusions To our knowledge this is the first study demonstrating effects of combined O3 and N deposition on N pools and the fate of anthropogenic N in semi-natural ecosystems. The results reveal that both air pollutants have the potential to profoundly alter element pools and ecosystem properties of subalpine grassland through an accumulation of N in either plant pools (N) or soil (O3). These findings are important with regard to the ecological consequences of increasing levels of the two important forms of regional air pollution. Although the accumulated N seems to be stabilized in a comparatively stable form in both cases, there is a risk that in the long run N is released to the environment once the conditions are changing. This underlines the importance of longterm field experiments with established vegetation providing the possibility to investigate ecological processes in situ under most realistic conditions. Acknowledgements This work was supported by the Swiss Federal Office for the Environment FOEN (A2310.0124) in the framework of the UNECE Convention on Long-Range Transboundary Air Pollution CLRTAP and the EU project ECLAIRE. The help of Anne-Lena Wahl, Ulrike Zell, and Robin Giger for performing field work and laboratory assistance is greatly acknowledged. Appendix A. Supplementary data Supplementary data related to this article can be found at http:// dx.doi.org/10.1016/j.envpol.2015.02.038. References Aerts, R., deCaluwe, H., 1997. Nutritional and plant-mediated controls on leaf litter decomposition of Carex species. Ecology 78, 244e260. Andersen, C.P., 2003. Source-sink balance and carbon allocation below ground in plants exposed to ozone. New Phytol. 157, 213e228. Arnone, J.A., Hirschel, G., 1997. Does fertilizer application alter the effects of elevated CO2 on Carex leaf litter quality and in situ decomposition in an alpine grassland? Acta Oecol. Int. J. Ecol. 18, 201e206.
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Elevated ozone and nitrogen deposition affect nitrogen pools of subalpine grassland €ch, Alain Valsangiacomo, Jochen Mayer, Hans-Rudolf Oberholzer, Seraina Bassin*, David Ka Matthias Volk, Jürg Fuhrer Agroscope, Institute of Sustainability Sciences ISS, Reckenholzstrasse 191, CH-8046 Zurich, Switzerland
a r t i c l e i n f o
a b s t r a c t
Article history: Received 5 December 2014 Received in revised form 25 February 2015 Accepted 27 February 2015 Available online
In a free-air fumigation experiment with subalpine grassland, we studied long-term effects of elevated ozone (O3) and nitrogen (N) deposition on ecosystem N pools and on the fate of anthropogenic N. At three times during the seventh year of exposure, N pools and recovery of a stable isotope tracer (15N) were determined in above- and belowground plant parts, and in the soil. Plants were much better competitors for 15N than soil microorganisms. Plant N pools increased by 30e40% after N addition, while soil pools remained unaffected, suggesting that most of the extra N was taken up and stored in plant biomass, thus preventing the ecosystem from acquiring characteristics of eutrophication. Elevated O3 caused an increase of N in microbial biomass and in stabilized soil N, probably resulting from increased litter input and lower litter quality. Different from individual effects, the interaction between the pollutants remained partly unexplained. © 2015 Published by Elsevier Ltd.
Keywords: Ozone Nitrogen Grassland Stable isotopes Soil
1. Introduction Concentrations of reactive nitrogen compounds (N) and tropospheric ozone (O3) have risen substantially since the industrial revolution (Galloway et al., 2004) and are expected to increase furthermore in the future (Ashmore, 2005). Both trace gases are known to threaten the biodiversity and functioning of ecosystems separately (Sutton et al., 2014), but large uncertainty exists about interactive effects between the two pollutants on (semi-)natural vegetation. Anthropogenic N input has been shown to affect plant productivity, community composition, and biodiversity especially in nutrient-poor semi-natural habitats (Bobbink et al., 2010). The impact of anthropogenic N on ecosystem functioning is modulated by the amount of N retained in soil and vegetation pools, whether it is stored, released to microorganisms and plants, or lost from soil via gaseous fluxes and leaching. In a comprehensive review including different temperate (semi-)natural ecosystems Phoenix et al. (2012) revealed both mineralization and nitrification to increase in five out of ten cases under elevated N deposition.
* Corresponding author. Agroscope, Climate/Air Pollution enholzstrasse 191, CH-8046 Zurich, Switzerland. E-mail address: [email protected] (S. Bassin). http://dx.doi.org/10.1016/j.envpol.2015.02.038 0269-7491/© 2015 Published by Elsevier Ltd.
Group
Reck-
Artic and alpine ecosystems are considered especially vulnerable to increasing N deposition since they have been exposed to very low deposition levels in the past. These soils are still nutrient poor and rich in organic matter due to the temperature limitation of € rner, 2003). As a consequence, the soil solution is mineralization (Ko dominated by dissolved organic N (DON) and traces of NHþ 4 , while free NO is typically missing (Chapin et al., 2012). In spring after 3 snowmelt, an N pulse is released from the decaying winter microbial biomass, providing a DON source for plants and microbes during the growing season (Schmidt et al., 2007). When mineral N is deposited into nutrient limited ecosystems, it is often instantly immobilized by soil microorganisms, since these are mostly C and N co-limited, especially in arctic and alpine soils (Chapin et al., 2012). €rner, With plants in these habitats being primarily N limited (Ko 2003), and in view of their capacity to take up and store excess N (Monson et al., 2006) the plant-microbe competition for N is €rner, 2003). Thus, some of the anthroconsidered to be severe (Ko pogenic N will accumulate in soil microorganisms and plants, reducing the risk for mineral N leaching for a certain time (Phoenix et al., 2012). However, remarkable differences exist among habitat types concerning the ecosystem pools that act as storage sites of the pollutant N: in lowland heath, two-thirds was retained in microbial biomass while in grasslands, bulk soil and plant pools were the dominant sinks (Phoenix et al., 2012). The stable isotope technique using a 15N enriched tracer is a widely used method to follow the
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fate of anthropogenic N in environmental systems (Dawson et al., 2002). In a Swiss mountain meadow, a substantial part of added tracer was instantly immobilized in soil, presumably bound to clay particles (Providoli et al., 2006), in line with results from a worldwide review on 15N tracer recoveries in temperate and arctic ecosystems (Templer et al., 2012). O3 is the most phytotoxic air pollutant in Europe (Ashmore, 2005). It primarily affects the carbon (C) budget of plants by reducing CO2 assimilation and by increasing the C investment into chemical and physical defense mechanisms (Andersen, 2003). But it has been shown to affect N budgets as well, e.g. by impairing N resorption from leaves before abscission, causing a substantial increase in whole-tree N loss through litter fall (Uddling et al., 2006). O3-induced changes in litter chemistry, which often include increased levels of phenolic compounds (Wang and Frei, 2011) may lead to reduced decomposition rates as shown in hardwood forests (Liu et al., 2009). In soybean, changes in litter decomposition processes were identified as cause for the increased levels of total N found in soil after six years of elevated O3 exposure (Pereira et al., 2011). Often, microbial activity was reduced by O3 fumigation as a result of reduced C allocation belowground (Andersen, 2003). While the direct impact of O3 on N budgets are expected to be small compared to those of increased N input, the magnitude of the indirect effects through reduced C availability is unknown. The interactive effects of elevated O3 and N deposition have been studied mainly in pot experiments with a focus on biomass production (e.g. Wyness et al., 2011) revealing N to alleviate the negative effects of O3 in most cases. Yet, long-term studies on N pools are missing in this context. Using a free air fumigation system we were able to expose intact monoliths of a species-rich subalpine grassland in situ at 2000 m a.s.l. to different combinations of O3 and N deposition for seven years. Substantial increases in aboveground biomass were recorded in response to N addition, mainly as a result of an extraordinary success of sedges, which tripled their relative abundance up to 30% (Bassin et al., 2013). In contrast, the O3 treatment did not affect above-and belowground biomass production (Volk et al., 2014) but it accelerated leaf senescence (Bassin et al., 2007b). The aim of the present study was to investigate the long-term effects of elevated O3 and N deposition, singly and in combination, on the soil and plant N pools of the subalpine pasture and the fate of anthropogenic N in this ecosystem. After a pulse application of a dual labelled 15NO15 3 NH4 tracer in May, soil cores were taken at three dates during the seventh and last growing season of the experiment and the ecosystem N pools (green phytomass, necromass, litter, roots, microbial biomass, soluble soil N, extractable soil N, stabilized soil N), their C:N ratio, as well as the recovery rates of the tracer was determined to test the following hypotheses: 1) The major fraction of added mineral N is bound to the mineral and organic fraction of the soil. 2) Elevated O3 causes accumulation of soil N in the form of soil organic matter (SOM), as a result of reduced microbial activity (due to lesser availability of labile C) in combination with less degradable litter. 3) Consequently, highest N accumulation in soil pools is expected when elevated O3 and N are combined.
2. Methods 2.1. Experiment The experiment was established at Alp Flix, a high plateau at 2000 m a.s.l. near Sur, Grisons, Switzerland (46 310 N/9 380 E). The
climate is characterized by a mean annual temperature of 1.1 C, a vegetation period lasting from April to October with average temperatures of 6.2 C and a permanent snow cover in winter (Volk et al., 2014). The annual sum of rainfall is approximately 1280 mm. The vegetation under investigation, a Geo-MontaniNardetum, is a typical perennial alpine pasture community covering large areas of the subalpine zone in the Alps and the Pyrenees. Grasses like Festuca rubra SCHLEICH. ex GAUD. and Nardus stricta L. and sedges like Carex sempervirens VILL. are the most abundant species (Bassin et al., 2013). The soil is a slightly acidic Cambisol with a depth varying between 20 and 40 cm, developed on a Serpentine bedrock. Rooting depth is 10e20 cm, with 80% of the roots found in the uppermost 7 cm. The experiment consisted of nine rings (Ø 7 m) of a free air O3 fumigation system with three different levels: ambient air (ca. 47 ppb); 1.35 ambient concentration, 1.73 ambient concentration, mean enrichment factor for the seven experimental years (for details see Volk et al., 2014). Rings were arranged at random in three blocks. Each ring contained 20 turf monoliths (40 cm long 30 cm wide 20 cm deep) excavated out of a nearby GeoMontani-Nardetum pasture, put into drained plastic boxes and randomly assigned to the rings. Five N levels equivalent to 0, 5, 10, 25, and 50 kg N ha1 y1 were applied in the form of ammonium nitrate dissolved in 200 ml of well water during the growing season every second week (solute concentration: up to 0.018 g N l1). Average wet and dry background deposition amounted to 4 kg N ha1 y1 (Bassin et al., 2007b). For the present study, we used a subsample of 36 monoliths exposed to either the highest concentration of O3 (1.73 ambient; “þO3”), the highest N-deposition (50 kg N ha1y1; ”þN00 ), or the combination of these two (“þN þ O3”), as well as from the untreated control (n ¼ 9 per treatment combination). 2.2. Sampling collection and processing The field work was accomplished in 2010, the seventh and last year of treatment. On 11 May, shortly after snowmelt, 0.04 g m2 of 15 4 N in the form of dual labelled 15NO15 3 NH (60 atom %) dissolved in 2.027 l water was sprayed onto the vegetation in the selected 36 monoliths, followed by the standard N application (see above). On 13 May, 12 July and 30 August, two cores each with a height of 10 cm and a diameter of 6 cm were taken in the monoliths. They were kept cool in boxes and were taken to the laboratory on the same day, where they were stored at 4 C until processing. Soil cores were halved and the halves were merged together to form two mixed samples of which one was sieved (2 mm mesh size) to get a root-free soil sample. Roots were removed with tweezers and finer roots that passed the sieve were recaptured during further processing of the soil (see below). The other sample was used to determine root biomass by washing on a sieve with 0.063 mm mesh size. Aboveground phytomass was cut at the soil surface of each core and green phytomass was separated from standing-dead (subsequently called “necromass”) and leaf litter (“litter”). On 12 July, aboveground biomass was harvested in two portions, i.e. above and below 2 cm, which was the standard cutting height used in the experiment for yield assessment (>2 cm ¼ “green phytomass removed through harvest”). All plant and soil samples were weighed after drying to constant weight at 60 C and then ground to a fine powder with a ball mill (Mixer Mill, Retsch). In the sieved fresh soil, microbial biomass C and N was determined using the chloroform fumigation extraction method (Brookes et al., 1985). Micro-roots were separated from the soil by a pre-extraction step (Olfs and Sherer, 1996). For this purpose 100 g fresh soil was extracted with 240 ml 0.05 M K2SO4 for 30 min. The
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interactions of them was made using likelihood ratio tests (Burnham and Anderson, 2002). Data were transformed when necessary. All analyses were carried out by means of the statistics software R (R Development Core Team, 2011) using the packages nlme, nortest, lattice, and car.
soil water suspension was collected in a beaker and left for 30 min to let the soil settle. The floating roots were sucked with a vacuum pump. Subsequently, the soil water suspension was filtered on a 589 paper filter (Schleicher and Schüll, Germany). The N obtained in this pre-extraction step was considered “soluble soil N”. The filtered soil was then divided into two halves. One halve was immediately extracted with 120 ml 0.5 M K2SO4 for 30 min (“extractable soil N”). The second half was fumigated with 25 ml chloroform in a desiccator for 24 h at 25 C and later extracted with 120 ml 0.5 M K2SO4. Total organic C (TOC) and total N (TNb) in soil extracts were determined by infrared spectrometry after combustion at 850 C (TOC) and subsequently by chemoluminescence (TNb) (DIMATOC 100, Dimatec, Essen, Germany). Microbial biomass C and N was calculated according to Joergensen (1996), using a kEN value of 0.54 (Joergensen and Mueller, 1996) and a kEC value of 0.45 (Joergensen, 1996). The 15N content of the solutions obtained in the three soil extraction steps was determined using a modified diffusion technique described by Mayer et al. (2003). The diffusion filters, the ground plant and soil samples were analysed for N and 15N content using a Flash EA1112 Series elemental analyzer (Thermo Italy Rhodano, Italy) coupled to a Finnigan MAT Delta plusXP isotope mass spectrometer (Finnigan MAT, Bremen, Germany). The 15Nisotope composition were expressed in standard notation (d15N) in parts per thousand (‰) relative to atmospheric N2 (standard). The d15N in the microbial biomass (MB) was calculated according to Mayer et al. (2003). All d 15N values were obtained by correcting the 15N enrichments with the background values of the N pools without labelling (green phytomass 2.8, necromass 2.6, roots 1.8, litter 2.4, soluble soil N 4.5, extractable soilunfum N 8.1, extractable soilfum N 3.1, stabilized soil N 2.7 d 15N). The 15N amount of the fumigated and extracted soil (“stabilized soil N”), soluble soil N (N extracted by 0.05 M K2SO4 solution), extractable N (N extracted by 0.5 M K2SO4 solution), and microbial biomass was calculated by multiplying its total N with the respective 15N abundance.
3. Results 3.1. Ecosystem properties Averaged over the growing season, in the control plots total ecosystem N pool was 364 g m2, of which the major part was stabilized in soil (317 g N m2) (Fig. 1a). The plant pools accounted for 5.8 g m2 aboveground, and 19.3 g m2 belowground. In addition, 18.3 g N m2 was stored in microbial biomass, while only 3.1 g m2 was soluble and extractable from soil. In July, approximately 1.8 g N m2 were allocated from roots to shoots, of which 1.2 ± 0.2 g m2 were removed through harvest (Fig. 1a), but by the end of the season, in August, N pools in living plant parts were almost equally distributed as in spring. N pools in necromass and litter decreased during the growing season by 11% and 65%, respectively (Figs. 1 and 2). 3.2. Tracer recovery Of the added stable isotope tracer, 67% were recovered two days after application in the control plots (Fig. 1b). 31 ± 4% were found in roots, 11 ± 1.1% were stabilized in soil, 12 ± 2% were found in necromass, 4.3 ± 0.7% in green phytomass, and 5 ± 1.4% in microbial biomass. In July, values decreased to 23 ± 4% in roots and increased to 14 ± 0.6% in green phytomass, of which 30% were exported through harvest. By the end of the season, 63% of the added tracer were recovered, 14.8 ± 1.7% were stabilized in soil, 25 ± 0.9% were found in roots, 9.1 ± 0.7% in necromass, 8.7 ± 0.7% in green phytomass, 4.3 ± 0.03% in microbial biomass and 0.6 ± 0.01% in litter. In most of the pools, the tracer recovery decreased significantly over time, while it increased in stabilized soil N (Fig. S1 and Table S2).
2.3. Statistical analyses 3.3. Nitrogen effects Effects of treatments on N pools and C:N ratios were tested in a repeated measures split-plot analysis using a linear mixed effect model (lme) (Pinheiro and Bates, 1996) with O3 at the main plot level (n ¼ 3) and N at the sub-plot level. All parameters were repeated by sampling date. Block, O3, and N entered the model as class variables, whereas time was a continuous variable. To account for the inherent variability among monoliths, initial grass cover (Grass2004), initial N content (Nsoil2003) and pH of the soil measured in 2003 before treatment onset were used as continuous co-variables. Inference on the significance of variables and selected
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Extra N addition increased the following N pools significantly compared to control plots (Fig. 2, þN and þN þ O3): green phytomass (þ31%) necromass (þ41%), litter (þ30%), roots (þ31%), and soluble soil N (þ22%). The C:N ratio was significantly reduced in green phytomass (12%), necromass (13%), roots (19%), and soluble N (17%) (Fig. 3), thus indicating that the extra N increased both the size as well as the N concentration of these pools by 20e30% and 10e17%, respectively (data not shown). In contrast, N pools and C:N ratio of microbial biomass, extractable soil N, and
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Fig. 1. a) Ecosystem N pools and b) the recovery of a stable isotope 15N tracer in subalpine grassland monoliths in the top 10 cm of untreated plots (control) at three dates during the growing season 2010 (13 May, 12 July, 30 August). The 15NH15 4 NO3 tracer was applied right after snowmelt on 11th May. At harvest, green phytomass above 2 cm was removed.
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Fig. 2. N pools of green phytomass, necromass, litter, roots, soluble N, extractable N, microbial biomass, and N stabilized in soil in plots exposed to elevated O3 concentrations (þO3), elevated N deposition (þN), the combination of them (þN þ O3), and in the control. Means ± 1SE (n ¼ 9) are displayed for three dates during the growing season 2010. Closed circles: control; open circles: þN; closed triangles: þO3; open triangles: þN þ O3. Significant factors and covariates (P < 0.05) inferred from likelihood ratio tests using linear mixed effect models are listed in the figures (for details see Table S1).
stabilized soil N remained unaffected. Only soluble N was increased by 22% throughout the growing season. N removed though harvest in þN and þN þ O3 treatments (2.6 ± 0.3 g m2) was twice the amount in the control (1.2 ± 0.2 g m2). However, below the cutting height, green phytomass still accounted for 2.47 ± 0.4 g N m2, but the N effect was only þ20%. Stable isotope recovery in N pools did not differ between Ntreated and control plots (Table S2), with three exceptions: compared to the control plots (5 ± 1.3%), exported green phytomass contained significantly more 15N (9 ± 1.6%) in N-treated plots (þN and þN þ O3), while 18% less 15N was stabilized in soil (Fig. S1). In the same monoliths, more tracer was recovered in soluble and extractable N, as well as in microbial biomass two days after tracer application compared to control plots, with the effect decreasing over time (significant N time interaction). 3.4. O3 effects Elevated O3 increased the N pool in green phytomass by 9.4% (¼
average of þO3 and þN þ O3 for the growing season, Fig. 2) compared to non-fumigated plots (þN and control) through a rise in N concentration (Fig. 3). Also, microbial biomass N pool was significantly increased by 9%, but at a constant C:N ratio (approximately 6.8), indicating that microbial biomass C was simultaneously stimulated by elevated O3. Remarkably, in response to elevated O3, N stabilized in soil increased significantly by 9% from 322 ± 7 g m2 in non-fumigated plots to 352 ± 3 g m2 (¼ average of þO3 and þN þ O3), but with no change in C:N ratio. Elevated O3 fumigation had no significant effect on 15N recovery in any of the ecosystem N pools (Table S2). 3.5. O3 N interactive effects Significant and consistent interactive O3 N effects were found in belowground N pools, namely in soluble soil N, and extractable soil N, with the effect of O3 being consistently negative in nonfertilized plots (þO3), but positive in fertilized treatments (þN þ O3, Fig. 2), compared to the control. For example, soluble soil
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Fig. 3. C:N ratio of green phytomass, necromass, litter, roots, soluble N, extractable N, microbial biomass, and N stabilized in soil in plots exposed to elevated O3 concentrations (þO3), elevated N deposition (þN), the combination of them (þN þ O3), and in control plots. Means ± 1SE (n ¼ 9) are displayed for three dates during the growing season 2010. Closed circles: control; open circles: þN; closed triangles: þO3; open triangles: þN þ O3. Significant factors and covariates (P < 0.05) inferred from likelihood ratio tests using linear mixed effect models are listed in the figures (for details see Table S3). Missing data for litter C content due to the low amount of litter available.
pools were 33% higher in þN þ O3 but 12% lower in þO3 compared to the control. Interactive effects of O3 and N were even more pronounced in the C:N ratio of roots (Fig. 3, Table S3), with values being 9% lower in þN, 9% higher in þO3, but 15% lower in þN þ O3 compared to the control, when averaged over the growing season. In microbial biomass C:N, the response pattern was similar and statistically significant (Fig. 3). 4. Discussion 4.1. Nitrogen effects N pools in plants, microbial biomass, soluble, and extractable N totalled 47.1 g N m2, similar to values for subalpine meadows (Robson et al., 2010) and alpine tundra (Fisk et al., 1998). Relative to this total pool size, the cumulative load of extra N (þ35 g m2 ¼ 7 5 g m2) applied during seven years in the þN treatment was very high, and sufficient to affect both the nutrient balance as well as ecosystem processes. The additional N was
primarily stored in plant pools, which gained not only in size due to increased biomass production, but also in N content through luxury uptake (Fig. 3) (Bassin et al., 2009), thus contrasting our first hypothesis. N capture for storage is common in alpine plants (Monson et al., 2006), as root biomass acts as a long-term reservoir for C and N (Van der Krift and Berendse, 2002). Here, much more N (up to 24 g N m2) was stored in roots compared to the maximum of 5.5 g N m2 found in green biomass and necromass, respectively, at peak vegetation development (Fig. 1). Also the yearly extra export of 1.4 g N m2 through harvest in þN plots is small, which limits the effectiveness of management options aiming to mitigate the impact of anthropogenic N deposition. Necromass played an important role as well, firstly due to its capacity to adsorb and store extra N (Fig. 1b, Fig. S1), and secondly because it showed the most pronounced enhancement in N pool (þ40%, Fig. 2) due to an accumulation of N-rich senescent leaves, related to the increase of sedges such as C. sempervirens (Bassin et al., 2013). Tough fibrous leaves of sedges are characterized by low decomposability (Aerts and deCaluwe, 1997), which is not improved through higher N contents after N addition (Arnone and Hirschel, 1997). Moreover, in
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tussocks of C. sempervirens decomposition rates are lowered by a drier microclimate (Yu et al., 2011). An increase in the abundance of Cyperaceae species in alpine vegetation is often observed in response to elevated N deposition (Bassin et al., 2012; Soudzilovskaia and Onipchenko, 2005). With the prevalence of poorly degrading graminoids, added nutrients are long-term stabilized in necromass, thus preventing the community from quickly acquiring features such as high decomposition rates and high nutrient availability typical of eutrophication (Soudzilovskaia et al., 2007). Plants were much better competitors for added N than microbes, as 48 h after the application of the tracer 35% of 15N was recovered in living plant parts while only 5% were found in microbial biomass. This contrasts results from studies revealing microorganisms as superior competitors for added mineral N (see the review of Kuzyakov and Xu (2013)). This discrepancy could be caused by the fact that in competition studies the tracer is typically injected directly into the rhizosphere, while here, in order to simulate N deposition from the atmosphere, the tracer was sprayed to the top of the canopy, which could have influenced the spatiotemporal pattern of N uptake. However in alpine tundra, plants were also the dominant N consumers, especially in spring when microbial biomass decreased after snowmelt as a result of C limitation (Lipson et al., 1999). It is likely that plants, regardless of their €sholm et al., 2009), prefer mineral capacity to take up organic N (Na N if available, while microbes preferably make use of excess DON in soil solution (Chapin et al., 2012). As a consequence, microbial biomass N pool and C:N ratio remained unaffected by N addition. Similarly, Manning et al. (2008) found no significant effect of N addition on microbial biomass, thus supporting the assumption that N limitation is not widespread among microbes (Treseder, 2008). The low C:N ratio (6.6:1 in control plots) in comparison to usually observed ratios of around 10:1 (Chapin et al., 2012) may explain the absence of N shortage in the present microbial community. Finally, despite the high N-sensitivity of this vegetation in terms of species composition and biodiversity (Bassin et al., 2013), the risk of N-induced acceleration of N cycling, enhanced N leaching or gaseous N losses to the atmosphere seems to be comparatively low. E.g. leachate collected during the growing season 2010 in a subsample of 20 monoliths revealed an accumulated N loss of only 0.15 g m2 (90% as DON, unpublished data), irrespective of N treatment. Evidently the N cycle in this ecosystem is still tightly closed even after seven years of extra N input. But it remains possible that N stored in plants may become mobilized in the long run due to external triggers such as climate warming or altered precipitation patterns, with unknown consequences for speciescomposition, N- and C fluxes, and N-leaching (Phoenix et al., 2012). 4.2. O3 effects In response to elevated O3, the N pool in green phytomass increased by approximately 9%, as compared to non-fumigated plots, due to higher leaf N concentration. In agreement, in a meta-analysis Wittig et al. (2009) found consistently increased N concentrations in O3 exposed tree leaves and explained this by retranslocation of N from prematurely senescing parts of the crown. Alternatively, N resorption from senescing leaves could also be reduced (Uddling et al., 2006) as a result of both impaired phloem loading and apoptosis-like cell death preventing complete degradation of organelles (Fukuda and Greenberg, 2000). Theoretically, the preferential shift in N allocation towards aboveground phytomass could negatively affect belowground N pools, but due to the low shoot:root ratio typical for these subalpine grasslands, the aboveground increase by 0.5 g N m2 would not lead to a detectable
change in the root N pool, which accounts for 14e24 g N m2 at harvest date. Equally, export of N-enriched phytomass through harvest would not cause noticeable changes in the ecosystem N budget. Similar to green phytomass, elevated O3 also increased the N pool in soil microbial biomass throughout the growing season. Similar results were found in O3-fumigated forests and boreal peat land, presumably as a result of altered litter quality (Morsky et al., 2008; Scagel and Andersen, 1997). Here, at elevated O3 a higher proportion of necromass was found in harvested phytomass in several consecutive years (Bassin et al., 2013), and presumably, soft leaves typical of O3-sensitive species (Bassin et al., 2007a) senesce earlier and are rapidly decomposed thus providing a readily available substrate for microorganisms. Preliminary results from the present experiment indicate higher amounts of soil microbial DNA in the O3-treated plots, however only in July, thus supporting the notion of increased substrate availability at the peak of vegetation development (Schneider, pers. communication). But, with root biomass being two to three times larger than aboveground plant € rner, 2003), roots biomass in alpine and arctic ecosystem (Ko remain the dominant C and nutrient input (Smith et al., 2014), and here the C:N ratio of microorganisms mirrored the ratio in response to the treatment combinations (Fig. 3). Thus, increased root turnover rates could be a more likely explanation for the O3-induced stimulation of microbial biomass. While many studies revealed biomass of roots to decrease under O3 fumigation, namely in fast growing establishing plants and crops (Grantz et al., 2006), only a few investigated root turnover rates. These were stimulated in mature O3-stressed trees (Haberer et al., 2007; Pregitzer et al., 2008) as a possible alternative strategy for nutrient acquisition when mycorrhizal infection rates decline (Wang et al., 2014) due to restricted allocation of soluble carbohydrates. However, due to the lack of supporting data, the mechanism leading to the observed increase in the microbial N pool remains open. As a further consequence of elevated O3 exposure, the pool of stabilized soil N increased significantly, in agreement with our second hypothesis, and the C:N ratio remained unchanged (Fig. 3) thus suggesting enhanced accumulation of poorly degradable soil organic matter. Increased root and leaf litter input combined with changed litter quality could be the cause. In soybean, increased levels of total soil N were found after six years of elevated O3 exposure as a consequence of altered litter decomposition processes (Pereira et al., 2011). A low ratio of soluble to structural carbohydrates (Kontunen-Soppela et al., 2007) as well as higher contents of phenolic compounds (Wang and Frei, 2011) are frequently observed physical defence mechanisms against O3 stress, leading to lower degradability of plant material (Booker et al., 2005). In addition, soil C and N turnover rates could have been slowed by the expansion of the slow-growing stress-tolerant N. stricta in þO3 and þN þ O3 plots (Bassin et al., 2013). According to Personeni et al. (2005) the rhizosphere priming effect through root exudates among grass species is dependent on their growth strategy, suggesting lower C and N turnover rates in less competitive species. In the long run, this O3-induced accumulation of SOM has the potential to substantially alter both structure and function of the ecosystem in terms of nutrient cycling, water retention capacity, but also C sequestration rates in a positive way, eventually counterbalancing the negative effects of O3 on C-sequestration expected from growth reductions (see Harmens and Mills, 2012). 4.3. O3 N effects We expected highest N accumulation in þN þ O3 plots as a consequence of reduced litter degradation in þO3 plots in
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combination with increased soil N storage in the þN treatment (third hypothesis). In fact, the largest extractable and soluble soil N pools were observed in this treatment combination (Fig. 2). Also, C:N ratio of roots and microbial biomass was reduced, suggesting highest N availability and quick N turnover rates in þN þ O3. However, the same was not observed for N stabilized in soil, for which the largest pool was found in þO3 (Fig. 2). As an explanation, in þN þ O3 plots N addition could counteract the effect of O3 on SOM accumulation through N-induced accelerated litter decomposition. However, evidence for this explanation remains weak, and because the 15N tracer study did not improve our understanding of the underlying processes (Fig. S1), it cannot be ruled out that besides the co-variables used in the model (soil pH, original soil N content, original grass cover) additional factors were involved in the observed N O3 interaction. Our data suggest that in accordance to aboveground biomass, individual species abundance and performance, the interactive effects of O3 and N on ecosystem N pools were ambiguous. Rather, the two pollutants acted individually: while the effects of N were limited to the plant N pools, O3-induced changes in C metabolism indirectly caused changes in soil C and N urnover processes. But, the interpretation of the results partly relies on assumptions which could not be validated due to a lack of comparable O3 N experiments. 5. Conclusions To our knowledge this is the first study demonstrating effects of combined O3 and N deposition on N pools and the fate of anthropogenic N in semi-natural ecosystems. The results reveal that both air pollutants have the potential to profoundly alter element pools and ecosystem properties of subalpine grassland through an accumulation of N in either plant pools (N) or soil (O3). These findings are important with regard to the ecological consequences of increasing levels of the two important forms of regional air pollution. Although the accumulated N seems to be stabilized in a comparatively stable form in both cases, there is a risk that in the long run N is released to the environment once the conditions are changing. This underlines the importance of longterm field experiments with established vegetation providing the possibility to investigate ecological processes in situ under most realistic conditions. Acknowledgements This work was supported by the Swiss Federal Office for the Environment FOEN (A2310.0124) in the framework of the UNECE Convention on Long-Range Transboundary Air Pollution CLRTAP and the EU project ECLAIRE. The help of Anne-Lena Wahl, Ulrike Zell, and Robin Giger for performing field work and laboratory assistance is greatly acknowledged. Appendix A. Supplementary data Supplementary data related to this article can be found at http:// dx.doi.org/10.1016/j.envpol.2015.02.038. References Aerts, R., deCaluwe, H., 1997. Nutritional and plant-mediated controls on leaf litter decomposition of Carex species. Ecology 78, 244e260. Andersen, C.P., 2003. Source-sink balance and carbon allocation below ground in plants exposed to ozone. New Phytol. 157, 213e228. Arnone, J.A., Hirschel, G., 1997. Does fertilizer application alter the effects of elevated CO2 on Carex leaf litter quality and in situ decomposition in an alpine grassland? Acta Oecol. Int. J. Ecol. 18, 201e206.
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